United
States
Office
of
Science
Environmental
Protection
and
Technology
June
30,
2002
Agency
Washington,
D.
C.
EPA­
822­
R­
03­
018
DRINKING
WATER
CRITERIA
DOCUMENT
ON
BROMINATED
TRIHALOMETHANES
DRAFT
February
20,
2003
Prepared
For
Health
and
Ecological
Criteria
Division
Office
of
Science
and
Technology
Office
of
Water
U.
S.
Environmental
Protection
Agency
Washington,
D.
C.
20460
Under
EPA
Contract
No.
68­
C­
99­
206
Work
assignment
0­
16
by
Syracuse
Research
Corporation
6225
Running
Ridge
Road
North
Syracuse,
NY
13212
Under
subcontract
to
The
Cadmus
Group,
Inc.
135
Beaver
Street
Waltham,
MA
02452
FOREWORD
Section
1412
(
b)
(
3)
(
A)
of
the
Safe
Drinking
Water
Act,
as
amended
in
1986
requires
the
Administrator
of
the
Environmental
Protection
Agency
to
publish
Maximum
Contaminant
Level
Goals
(
MCLGs)
and
promulgate
National
Primary
Drinking
Water
Regulations
for
each
contaminant,
which,
in
the
judgment
of
the
Administrator,
may
have
an
adverse
effect
on
public
health
and
which
is
known
or
anticipated
to
occur
in
public
water
systems.
The
MCLG
is
nonenforceable
and
is
set
at
a
level
at
which
no
known
or
anticipated
adverse
health
effects
in
humans
occur
and
which
allows
for
an
adequate
margin
of
safety.
Factors
considered
in
setting
the
MCLG
include
health
effects
data
and
sources
of
exposure
other
than
drinking
water.

This
document
provides
the
health
effects
basis
to
be
considered
in
establishing
the
MCLGs
for
brominated
trihalomethanes
found
in
chlorinated
drinking
water.
To
achieve
this
objective,
data
on
pharmacokinetics,
human
exposure,
acute
and
chronic
toxicity
to
animals
and
humans,
epidemiology
and
mechanisms
of
toxicity
were
evaluated.
Specific
emphasis
is
placed
on
literature
data
providing
dose­
response
information.
Thus,
while
the
literature
search
and
evaluation
performed
in
support
of
this
document
was
comprehensive,
only
the
reports
considered
most
pertinent
in
the
derivation
of
the
MCLGs
are
cited
in
this
document.
The
comprehensive
literature
search
in
support
of
this
document
includes
information
published
up
to
July
2001,
however,
more
recent
information
may
have
been
added
during
the
review
process.

When
adequate
health
effects
data
exist,
Health
Advisory
values
for
less
than
lifetime
exposure
(
One­
day,
Ten­
day
and
Longer­
term,
approximately
10%
of
an
individual's
lifetime)
are
included
in
this
document.
These
values
are
not
used
in
setting
the
MCLGs,
but
serve
as
informal
guidance
to
municipalities
and
other
organizations
when
emergency
spills
or
contamination
situations
occur.

Geoffrey
Grubbs
Director,
Office
of
Science
and
Technology
Office
of
Water
Acknowledgments
This
document
is
derived
and
updated/
expanded
of
the
Draft
for
the
Drinking
Water
Criteria
Document
on
Trihalomethanes
(
U.
S.
EPA,
1994)
and
the
Summary
of
New
Health
Effects
Data
on
Drinking
Water
Disinfectants
and
Disinfectant
Byproduct
(
D/
DBPs)
for
the
Notice
of
Availability
(
NODA)
(
U.
S.
EPA,
1997).
This
document
includes
an
evaluation
of
literature
on
Brominated
Trihalomethanes
resulting
from
a
full
literature
search
for
toxicity
data
conducted
in
July,
2001.
In
addition,
few
newer
studies
identified
after
the
literature
search
date
have
been
included
as
available
at
the
time
of
document
preparation.

Chemical
Manager:

Nancy
Chiu,
Ph.
D.

External
Peer
Reviewers:

Annette
L.
Bunge,
Ph.
D.
Colorado
School
of
Mines
John
Reif,
M.
Sc.
(
Med),
D.
V.
M.,
Colorado
State
university
Judy
Buelke­
Sam,
M.
A.
Toxicology
Services
TABLE
OF
CONTENTS
I.
EXECUTIVE
SUMMARY
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I
­
1
II.
PHYSICAL
AND
CHEMICAL
PROPERTIES
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II
­
1
A.
Properties
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II
­
1
B.
Summary
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II
­
2
III.
TOXICOKINETICS
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III
­
1
A.
Absorption
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III
­
1
B.
Distribution
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III
­
3
C.
Metabolism
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III
­
5
D.
Excretion
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III
­
13
E.
Bioaccumulation
and
Retention
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III
­
13
F.
Summary
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III
­
13
IV.
HUMAN
EXPOSURE
.
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IV
­
1
A.
Occurrence
in
Drinking
Water
.
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.
IV
­
1
1.
National
Surveys
.
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IV
­
2
2.
Other
Studies
.
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IV
­
8
3.
Estimates
of
Tap
Water
Ingestion
Exposure
to
Brominated
Trihalomethanes
.
.
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.
IV
­
13
B.
Exposure
from
Sources
Other
Than
Drinking
Water
.
.
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.
IV
­
17
1.
Dietary
Intake
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IV
­
17
2.
Air
Intake
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IV
­
20
3.
Concentrations
and
Exposures
Associated
with
Swimming
Pools
and
Hot
Tubs
.
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IV
­
27
4.
Soil
Concentrations
and
Exposure
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IV
­
30
C.
Overall
Exposure
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IV
­
30
D.
Body
Burden
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IV
­
31
1.
Blood
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IV
­
31
2.
Mother's
Milk
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IV
­
35
E.
Summary
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IV
­
35
V.
HEALTH
EFFECTS
IN
ANIMALS
.
.
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V
­
1
A.
Acute
Exposures
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V
­
1
1.
Bromodichloromethane
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V
­
1
2.
Dibromochloromethane
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V
­
6
3.
Bromoform
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V
­
8
B.
Short­
Term
Exposures
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V
­
8
1.
Bromodichloromethane
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V
­
14
2.
Dibromochloromethane
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V
­
21
3.
Bromoform
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V
­
24
C.
Subchronic
Exposure
.
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V
­
27
1.
Bromodichloromethane
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V
­
30
2.
Dibromochloromethane
.
.
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V
­
31
3.
Bromoform
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V
­
33
D.
Chronic
Exposure
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V
­
34
1.
Bromodichloromethane
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.
V
­
34
2.
Dibromochloromethane
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V
­
38
3.
Bromoform
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V
­
39
E.
Reproductive
and
Developmental
Effects
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.
.
.
.
.
.
.
.
.
.
.
V
­
41
1.
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
41
2.
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
56
3.
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
59
F.
Mutagenicity
and
Genotoxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
67
1.
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
67
2.
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
74
3.
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
81
G.
Carcinogenicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
85
1.
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
85
2.
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
92
3.
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
95
H.
Other
Key
Health
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
97
1.
Immunotoxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
97
2.
Hormonal
disruption
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
99
3.
Structure­
Activity
Relationships
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
100
I.
Summary
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
101
1.
Health
Effects
of
Acute
and
Short
Term
Exposure
of
Animals
.
.
.
.
V
­
101
2.
Health
Effects
of
Longer­
term
Exposure
of
Animals
.
.
.
.
.
.
.
.
.
.
.
V
­
101
3.
Reproductive
and
Developmental
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
102
4.
Mutagenicity
and
Genotoxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
103
5.
Carcinogenicity
Studies
in
Animals
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
103
6.
Other
Key
effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
104
VI.
HEALTH
EFFECTS
IN
HUMANS
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
1
A.
Clinical
Case
Studies
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
1
1.
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
1
2.
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
1
3.
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
1
B.
Epidemiological
Studies
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
1
1.
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
7
2.
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
11
3.
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
15
C.
High
Risk
Populations
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
15
D.
Summary
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
16
VII.
MECHANISM
OF
TOXICITY
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
1
A.
Role
of
Metabolism
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
1
B.
Biochemical
Basis
of
Toxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
1
C.
Mode
of
Action
of
Carcinogenesis
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
2
D.
Interactions
and
Susceptibilities
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
5
1.
Potential
Interactions
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
5
2.
Greater
Childhood
Susceptibility
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
6
3.
Other
Potentially
Susceptible
Populations
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
13
E.
Summary
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VII
­
16
VIII.
QUANTIFICATION
OF
TOXICOLOGICAL
EFFECTS
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
1
A.
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
1
1.
Noncarcinogenic
effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
1
2.
Carcinogenic
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
28
B.
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
35
1.
Noncarcinogenic
effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
35
2.
Carcinogenic
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
49
C.
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
55
1.
Noncarcinogenic
effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
55
2.
Carcinogenic
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
70
D.
Summary
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
75
IX.
REFERENCES
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IX
­
1
APPENDIX
A
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
A
­
1
APPENDIX
B
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
B
­
1
APPENDIX
C
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
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.
.
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.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
C
­
1
LIST
OF
FIGURES
Figure
III­
1
Proposed
Metabolic
Pathways
for
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
.
.
.
III
­
6
Figure
V­
2
Proposed
Routes
for
GST­
Mediated
Metabolic
Activation
of
Trihalomethanes
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
70
LIST
OF
TABLES
Table
I­
1
Summary
of
Quantification
of
Toxicological
Effects
for
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
.
.
.
.
.
I
­
15
Table
II­
1
Physical
and
Chemical
Properties
of
the
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
II
­
1
Table
III­
1
Recovery
of
Label
8
Hours
after
Oral
Administration
of
14C­
Labeled
Brominated
Trihalomethanes
to
Male
Sprague­
Dawley
Rats
or
Male
B6C3F
1
Mice
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
III
­
1
Table
III­
2
Cumulative
Excretion
of
Label
after
Oral
Administration
of
14C­
Labeled
Bromodichloromethane
to
Male
F344
Rats
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
III
­
2
Table
III­
3
Over
view
of
Tissue
Collection
for
Analysis
of
Bromodichloromethane
in
Sprague­
Dawley
Rat
Tissues
and
Fluids
(
CCC,
2000c)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
III
­
4
Table
IV­
1.
Brominated
Trihalomethane
Concentrations
Measured
in
U.
S.
Public
Drinking
Water
Systems
Serving
100,000
or
More
Persons
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
7
Table
IV­
2
NRWA
Brominated
Trihalomethane
Results
for
Small
Surface
Water
Plants
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
8
Table
IV­
3
Bromodichloromethane
Concentrations
in
Drinking
Water
from
the
U.
S.
EPA
TEAM
Study
(
µ
g/
L)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
9
Table
IV­
4
Dibromochloromethane
Concentrations
in
Drinking
Water
from
the
U.
S.
EPA
TEAM
Study
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
10
Table
IV­
5
Bromoform
Concentrations
in
Drinking
Water
from
the
U.
S.
EPA
TEAM
Study
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
11
Table
IV­
6
Estimated
Drinking
Water
Exposures
to
Brominated
Trihalomethanes
in
U.
S.
Public
Drinking
Water
Systems
Serving
More
than
100,000
Persons
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
14
Table
IV­
7.
Estimated
Distribution
of
Drinking
Water
Exposures
to
Brominated
Trihalomethanes
for
Populations
in
U.
S.
EPA
TEAM
Study
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
16
Table
IV­
8.
Selected
Concentration
Data
for
Individual
Brominated
Trihalomethanes
(
ppt)
in
Outdoor
Air
as
Summarized
in
Brodzinsky
and
Singh
(
1983)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
22
Table
IV­
9
Mean
Bromodichloromethane
Concentrations
in
Blood
Following
Three
Types
of
Water
Use
Events
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
33
Table
IV­
10
Median
Tap
Water
Trihalomethane
Levels
(
ppb)
in
Cobb
County
and
Corpus
Christi
Homes,
Water
Treatment
Plants,
and
Distribution
Systems
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
34
Table
IV­
11
Between
Site
Comparison
of
Median
Blood
Levels
(
ppt)
and
Changes
in
Blood
Levels
(
ppt)
after
Showering
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
IV
­
35
Table
V­
1
Summary
of
LD
50
Values
for
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
1
Table
V­
2
Summary
of
Acute
Toxicity
Studies
for
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
.
V
­
2
Table
V­
3
Summary
of
Short
Term
Toxicity
Studies
for
Brominated
Trihalomethanes
.
.
.
.
V
­
9
Table
V­
4
Summary
of
Subchronic
Toxicity
Studies
for
Brominated
Trihalomethanes
.
.
.
V
­
28
Table
V­
5
Summary
of
Chronic
Toxicity
Studies
for
Brominated
Trihalomethanes
.
.
.
.
.
.
V
­
35
Table
V­
6
NTP
(
1998)
Study
Design
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
44
Table
V­
7
Summary
of
Experiments
Conducted
by
Bielmeier
et
al.
(
2001)
.
.
.
.
.
.
.
.
.
.
.
.
V
­
46
Table
V­
8
Mean
Consumed
Doses
(
mg/
kg­
day)
of
Bromodichloromethane
in
the
Range
Finding
Study
Conducted
by
CCC
(
2000c)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
49
Table
V­
9
Summary
of
Reproductive
Studies
of
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
.
V
­
61
Table
V­
10
Summary
of
Mutagenicity,
Genotoxicity,
and
Neoplastic
Transformation
Data
for
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
72
Table
V­
11
Summary
of
Mutagenicity,
Genotoxicity,
and
Neoplastic
Transformation
Data
for
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
79
Table
V­
12
Summary
of
Mutagenicity,
Genotoxicity,
and
Neoplastic
Transformation
Data
for
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
84
Table
V­
13
Tumor
Frequencies
in
F344/
N
Rats
and
B6C3F
1
Mice
Exposed
to
Bromodichloromethane
in
Corn
Oil
for
2
Years
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
86
Table
V­
14
Hepatic
and
Renal
Tumors
in
Male
F344/
N
Rats
Administered
Bromodichloromethane
in
the
Drinking
Water
for
Two
Years
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
90
Table
V­
15
Frequencies
of
Liver
Tumors
in
B6C3F
1
Mice
Administered
Dibromochloromethane
in
Corn
Oil
for
105
Weeks
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
94
Table
V­
16
Tumor
Frequencies
in
the
Large
Intestine
of
F344/
N
Rats
Exposed
to
Bromoform
in
Corn
Oil
for
2
Years
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
V
­
95
Table
VI­
1
Epidemiological
Studies
Investigating
an
Association
Between
Chlorinated
Drinking
Water
and
Cancer
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
2
Table
VI­
2
Epidemiological
Studies
Investigating
an
Association
Between
Chlorinated
Drinking
Water
and
Adverse
Pregnancy
Outcomes
or
Altered
Menstrual
Function
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
3
Table
VI­
3
Means
and
Adjusted
Differences
in
Menstrual
Cycle
and
Follicular
Phase
Length
by
Quartile
of
Individual
and
Summed
Brominated
Trihalomethanes
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VI
­
13
Table
VIII­
1
Summary
of
Candidate
Studies
for
Derivation
of
the
One­
day
HA
for
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
4
Table
VIII­
2
Summary
of
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
10
Table
VIII­
3
Summary
of
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
for
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
17
Table
VIII­
4
Summary
of
Candidate
Studies
for
Derivation
of
the
RfD
for
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
22
Table
VIII­
5
Summary
of
Preliminary
BMD
Modeling
Results
for
the
Bromodichloromethane
RfD
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
25
Table
VIII­
6
Tumor
Frequencies
in
Rats
and
Mice
Exposed
to
Bromodichloromethane
in
Corn
Oil
for
2
Years
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
32
Table
VIII­
7
Summary
of
Cancer
Risk
Estimates
for
Bromodichloromethane
.
.
.
.
.
.
.
.
VIII
­
34
Table
VIII­
8
Summary
of
Candidate
Studies
for
Derivation
of
the
One­
day
HA
for
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
36
Table
VIII­
9
Summary
of
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
37
Table
VIII­
10
Summary
of
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
for
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
41
Table
VIII­
11
Summary
of
Candidate
Studies
for
Derivation
of
the
RfD
for
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
45
Table
VIII­
12
Results
of
Preliminary
BMD
Modeling
of
Selected
Data
from
NTP
(
1985)
Studies
VIII
­
47
Table
VIII­
13
Frequencies
of
Liver
Tumors
in
Mice
Administered
Dibromochloromethane
in
Corn
Oil
for
105
Weeks
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
VIII
­
53
Table
VIII­
15
Summary
of
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
.
.
VIII
­
57
Table
VIII­
16
Summary
of
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
for
Bromoform
.
.
.
.
.
.
.
.
.
.
.
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.
.
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.
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.
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.
.
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.
.
.
.
.
.
.
VIII
­
60
Table
VIII­
17
Summary
of
Candidate
Studies
for
Derivation
of
the
RfD
for
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
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.
.
.
.
.
.
.
VIII
­
66
Table
VIII­
18
Tumor
Frequencies
in
Rats
Exposed
to
Bromoform
in
Corn
Oil
for
2
Years
.
.
.
.
.
.
.
.
.
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.
.
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.
.
.
VIII
­
73
Table
VIII­
19
Carcinogenic
Risk
Estimates
for
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII
­
74
Table
VIII­
20
Summary
of
Advisory
Values
for
Bromodichloromethane,
Dibromochloromethane,
and
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
VIII
­
75
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
.
.
.
.
.
A
­
3
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dibromochloromethane
.
.
.
.
A
­
15
Table
A­
3
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromoform.
.
.
.
.
.
.
.
.
.
.
.
.
A
­
24
Table
A­
4
Model
Equations
used
in
BMD
Calculations
for
Health
Advisories
.
.
.
.
.
.
.
.
.
.
A
­
29
Table
A­
5
Benchmark
Dose
Modeling
Results
for
Bromodichloromethane
.
.
.
.
.
.
.
.
.
.
.
.
A
­
41
Table
A­
6
Benchmark
Dose
Modeling
Results
for
Dibromochloromethane
.
.
.
.
.
.
.
.
.
.
.
.
A
­
50
Table
A­
7
Benchmark
Dose
Modeling
Results
for
Bromoform
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
A
­
56
Table
C­
1
DBCM
Concentrations
Measured
in
U.
S.
Public
Drinking
Water
Systems
Serving
100,000
or
More
Persons
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
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.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
C
­
4
Table
C­
2
Selected
Concentration
Data
for
Individual
Brominated
Trihalomethanes
(
ppt)
in
Outdoor
Air
as
Summarized
in
Brodzinsky
and
Singh
(
1983)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
C
­
5
Table
C­
3
Results
of
RSC
Calculations
for
DBCM
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
C
­
17
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
1
I.
EXECUTIVE
SUMMARY
Brominated
trihalomethanes
are
volatile
organic
liquids
that
have
a
number
of
industrial
and
chemical
uses.
The
chief
reason
for
health
concern
is
that
they
are
generated
as
by­
products
during
the
disinfection
of
drinking
water.
The
brominated
trihalomethanes
occurring
in
water
are
bromoform,
dibromochloromethane,
and
bromodichloromethane.
These
compounds
are
formed
when
hypochlorous
acid
oxidizes
any
bromide
ion
present
in
water
to
form
hypobromous
acid,
which
subsequently
reacts
with
organic
material
to
form
the
brominated
trihalomethanes.

Toxicokinetics
No
human
data
on
absorption
of
brominated
trihalomethanes
are
available.
Measurements
in
mice
and
rats
indicate
that
gastrointestinal
absorption
of
brominated
trihalomethanes
is
rapid
(
peak
levels
attained
less
than
an
hour
after
administration
of
a
gavage
dose)
and
extensive
(
63%
to
93%).
Most
studies
of
brominated
trihalomethane
absorption
have
used
oil­
based
vehicles.
A
study
in
rats
found
that
the
initial
absorption
rate
of
bromodichloromethane
was
higher
when
the
compound
was
administered
in
an
aqueous
vehicle
than
when
administered
in
a
corn
oil
vehicle.

Data
for
distribution
of
brominated
trihalomethanes
in
human
organs
and
tissues
are
limited.
Bromoform
was
found
primarily
in
the
stomach
and
lungs
of
a
human
overdose
victim,
with
lower
levels
detected
in
intestine,
liver,
kidney
and
brain.
Dibromochloromethane
was
found
in
1
of
42
samples
of
human
breast
milk
collected
from
women
living
in
urban
areas.
Radiolabeled
brominated
trihalomethanes
were
detected
in
a
variety
of
tissues
following
oral
dosing
in
rats
and
mice.
Approximately
1
to
4%
of
the
administered
dose
was
recovered
in
body
tissues
when
analysis
was
conducted
8
or
24
hours
post­
treatment.
The
highest
concentrations
were
detected
in
stomach,
liver,
blood,
and
kidneys
when
assayed
8
hours
after
administration
of
the
compounds.
Bromodichloromethane
was
detected
at
a
concentration
of
0.38
µ
g/
g
in
the
milk
of
one
of
three
female
rats
exposed
to
approximately
112
mg/
kg­
day
during
a
reproductive/
developmental
study.
Bromodichloromethane
was
not
detected
in
placentas,
amniotic
fluid,
or
fetal
tissue
collected
on
gestation
day
21
from
rats
exposed
to
doses
up
to
approximately
112
mg/
kg­
day
or
in
plasma
collected
from
postpartum
day
29
weanling
pups.
Bromodichloromethane
was
detected
at
concentrations
slightly
above
the
limit
of
detection
in
placentas
from
two
litters
born
to
rabbits
exposed
to
76
mg/
kg­
day.
Bromodichloromethane
was
detected
in
one
fetus
from
a
rabbit
exposed
to
76
mg/
kg­
day
"...
at
a
level
below
the
limit
of
detection".
Bromodichloromethane
was
not
detected
in
placentas
from
female
rabbits
exposed
to
doses
of
approximately
32
mg/
kg­
day,
or
in
amniotic
fluid
or
the
remaining
fetuses
from
rabbits
exposed
to
doses
of
approximately
76
mg/
kg­
day.

Brominated
trihalomethanes
are
extensively
metabolized
by
animals.
Metabolism
of
brominated
trihalomethanes
occurs
via
two
pathways.
One
pathway
predominates
in
the
presence
of
oxygen
(
the
oxidative
pathway)
and
the
other
predominates
under
conditions
of
low
oxygen
tension
(
the
reductive
pathway).
In
the
presence
of
oxygen,
the
initial
reaction
product
is
trihalomethanol
(
CX
3
OH),
which
spontaneously
decomposes
to
yield
the
corresponding
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
2
dihalocarbonyl
(
CX
2
O).
The
dihalocarbonyl
species
are
reactive
and
may
form
adducts
with
cellular
molecules.
When
intracellular
oxygen
levels
are
low,
the
trihalomethane
is
metabolized
via
the
reductive
pathway,
resulting
in
a
highly
reactive
dihalomethyl
radical
(°
CHX
2),
which
may
also
form
covalent
adducts
with
cellular
molecules.
The
metabolism
of
brominated
trihalomethanes
and
chloroform
appear
to
occur
via
the
same
pathways,
although
in
vitro
and
in
vivo
data
suggest
that
metabolism
via
the
reductive
pathway
occurs
more
readily
for
brominated
trihalomethanes.
Both
oxidative
metabolism
and
reductive
metabolism
of
trihalomethanes
appear
to
be
mediated
by
cytochrome
P450
isoforms.
The
identity
of
cytochrome
P450
isoforms
that
metabolize
brominated
trihalomethanes
has
been
investigated
in
several
studies
which
used
bromodichloromethane
as
a
substrate.
The
available
data
suggest
that
the
cytochrome
P450
isoforms
CYP2E1,
CYP2B1/
2,
and
CYP1A2
metabolize
bromodichloromethane
in
rats.
The
human
isoforms
CYP2E1,
CYP1A2,
and
CYP3A4
show
substantial
activity
toward
bromodichloromethane
in
vitro
and
low
but
measurable
levels
of
CYP2A6
activity
have
also
been
detected.
Based
on
the
available
data,
CYP2E1
and
CYP1A2
are
the
only
isoforms
active
in
both
rats
and
humans.
CYP2E1
shows
the
highest
affinity
for
bromodichloromethane
in
both
species
and
the
metabolic
parameters
K
m
and
k
cat
are
similar
for
rat
and
human
CYP2E1.
In
contrast,
the
metabolic
parameters
for
CYP1A2
differ
in
rats
and
humans.
The
pattern
of
results
for
isozyme
activity
obtained
from
an
inhalation
study
of
bromodichloromethane
was
similar
to
the
pattern
reported
for
male
F344
rats
treated
with
bromodichloromethane
by
gavage.

The
lung
is
the
principle
route
of
excretion
in
rats
and
mice.
Studies
with
14C­
labeled
compounds
indicate
that
up
to
88%
of
the
administered
dose
can
be
found
in
exhaled
air
as
carbon
dioxide,
carbon
monoxide,
and
parent
compound.
Excretion
in
the
urine
generally
appears
to
be
5%
or
less
of
the
administered
oral
dose.
Data
from
one
study
suggest
that
fecal
excretion
accounts
for
less
than
3%
of
the
administered
dose.

Human
Exposure
Brominated
trihalomethanes
are
found
in
virtually
all
water
treated
for
drinking;
however,
concentrations
of
individual
forms
vary
widely
depending
on
the
type
of
water
treatment,
locale,
time
of
year,
sampling
point
in
the
distribution
system,
and
source
of
the
drinking
water.
Occurrence
data
for
brominated
trihalomethanes
are
available
from
13
national
surveys
and
9
additional
studies
that
are
more
restricted
in
scope.
The
procedures
used
for
sampling
processing
and
storage
and
calculation
of
summary
statistics
should
be
carefully
considered
when
evaluating
and
comparing
brominated
trihalomethane
occurrence
data.
Some
methods
restrict
trihalomethane
formation
by
refrigeration
or
the
use
of
quenching
agents,
whereas
others
maximize
trihalomethane
formation
by
storage
at
room
temperature.
Approaches
to
data
summarization
vary
in
their
treatment
of
data
below
the
analytical
detection
level
or
minimum
reporting
level.

When
all
available
national
survey
data
are
considered,
bromodichloromethane
concentrations
in
drinking
water
range
from
below
the
detection
limit
to
183
µ
g/
L
(
ppb),
while
dibromochloromethane
and
bromoform
concentrations
range
from
below
the
detection
limit
to
280
µ
g/
L
(
ppb).
When
data
for
the
three
brominated
trihalomethanes
are
compared,
the
frequency
of
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
3
detection
and
measured
concentrations
of
bromodichloromethane
in
drinking
water
supplies
tend
to
be
higher
than
those
for
dibromochloromethane.
Bromoform
is
detected
less
frequently
and
at
lower
concentrations
than
the
other
two
brominated
trihalomethanes,
except
in
some
ground
waters.
Concentrations
of
all
trihalomethanes
in
drinking
water
were
generally
lower
when
the
raw
water
is
obtained
from
ground
water
sources
rather
than
surface
water
sources.
The
most
recent
national
survey
data
are
those
collected
by
the
U.
S.
EPA
under
the
Information
Collection
Rule
(
ICR).
Monitoring
data
were
collected
over
an
18­
month
period
between
July
1997
and
December
1998
from
approximately
300
water
systems
operating
501
plants
and
serving
at
least
100,000
people.
Summary
occurrence
data
stratified
by
raw
water
source
(
groundwater
or
surface
water)
are
available
for
finished
water,
the
distribution
system
(
DS)
average,
and
the
DS
high
values.
The
mean,
median,
and
90th
percentile
values
for
surface
water
DS
average
concentrations
in
the
ICR
survey
are
8.6,
70.2,
and
20.3
µ
g/
L,
respectively,
for
bromodichloromethane
(
range
of
individual
values
0
­
65.8
µ
g/
L);
2.4,
4.72,
and
13.2
µ
g/
L,
respectively,
for
dibromochloromethane
(
range
0
­
67.3);
and
0.
1.18,
and
3.10,
respectively,
for
bromoform
(
range
0
­
3.43).

Exposure
to
brominated
trihalomethanes
via
ingestion
of
drinking
water
was
estimated
using
data
obtained
for
disinfectants
and
disinfection
byproducts
under
the
Information
Collection
Rule
(
ICR).
ICR
data
offer
several
advantages
over
other
national
studies
for
purposes
of
estimating
national
exposure
levels
of
adults
in
the
United
States
to
brominated
trihalomethanes
via
ingestion
of
drinking
water.
First,
they
are
recent
and
reflect
relatively
current
conditions.
Second,
data
of
very
similar
quality
and
quantity
were
collected
systematically
from
a
large
number
of
plants
(
501)
and
systems
(
approximately
300),
including
both
surface
and
ground
water
systems.
Third,
the
mean,
median,
and
90th
percentile
value
were
estimated
on
the
basis
of
all
samples
taken,
not
just
the
sample
detects.
Thus,
these
descriptive
statistics
are
representative
of
the
exposures
of
the
entire
populations
served
by
those
systems,
not
just
the
populations
served
by
systems
with
higher
concentrations
of
these
compounds.
However,
this
study
can
not
be
considered
representative
of
smaller
public
water
supplies
or
water
supplies
from
the
most
highly
industrialized
or
contaminated
areas.

Exposure
was
calculated
by
multiplying
the
concentration
of
individual
brominated
trihalomethanes
in
drinking
water
by
the
average
daily
intake,
assuming
that
each
individual
consumes
two
liters
of
water
per
day.
The
annual
median,
mean,
and
upper
90th
percentile
values
are
presented
for
both
surface
and
ground
water
systems.
Assuming
that
the
DS
High
value
actually
represents
the
average
exposure
level
of
persons
served
by
one
plant
distribution
pipe
with
the
longest
water­
residence
time,
the
DS
High
value
might
be
used
to
estimate
a
high­
end
exposure
level.

For
bromodichloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
17,
20,
and
40
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromodichloromethane
from
ground
water
systems
are
lower
 
3.6,
8.1,
and
22
µ
g/
person/
day,
respectively.
For
dibromochloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
4.8,
9.4,
and
26
µ
g/
person/
day,
respectively.
The
corresponding
values
for
populations
exposed
Draft
­
Do
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or
Quote
February
20,
2003
I
­
4
to
dibromochloromethane
from
groundwater
system
are
lower,
with
estimates
of
2.7,
6.2,
and
18
µ
g/
person/
day,
respectively.
For
bromoform,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
near
0,
2.4,
and
6.2
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromoform
from
ground
water
systems
are
higher,
with
estimates
of
0.65,
3.8,
and
9.6
µ
g/
person/
day,
respectively.

For
purposes
of
comparison,
estimates
of
ingestion
exposure
to
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
drinking
water
were
also
estimated
from
data
collected
in
other,
older
studies.
Ingestion
from
ground
water
supplies
was
estimated
from
the
median
levels
found
in
the
Ground
Water
Supply
Survey
conducted
by
U.
S.
EPA
in
1980­
81.
Based
on
the
range
of
median
levels
(
1.4
 
2.1
µ
g/
L
(
ppb))
and
a
consumption
rate
of
two
liters
per
day,
the
median
ingestion
exposure
to
bromodichloromethane
may
range
from
2.8
to
4.2
µ
g/
day.
Similarly,
median
exposure
to
dibromochloromethane
may
range
from
4.2
to
7.8
µ
g/
day,
and
for
bromoform,
median
exposure
may
range
from
4.8
to
8.4
µ
g/
day.
Exposure
to
bromodichloromethane
from
surface
water
supplies
can
be
estimated
based
on
the
range
of
median
values
observed
under
different
conditions
in
the
National
Organics
Monitoring
Survey
conducted
by
U.
S.
EPA
in
1976­
1977,
which
mainly
sampled
surface
water
systems.
Based
on
a
range
of
5.9
 
14
µ
g/
L
(
ppb),
exposure
to
bromodichloromethane
from
surface
water
is
estimated
to
be
between
12
and
28
µ
g/
day.
Similarly,
based
on
the
range
of
medians
reported
for
dibromochloromethane
concentrations,
the
median
exposure
is
estimated
to
be
up
to
6
µ
g/
day.
The
median
levels
of
bromoform
in
the
surface
water
supplies
have
been
found
to
be
less
than
the
EPA
Drinking
Water
minimum
reporting
levels
(
MRLs)
of
0.5
 
1
µ
g/
L
(
ppb).
An
estimate
of
exposure
based
on
the
MRLs
will
be
overly
conservative
because
the
actual
concentration
of
bromoform
is
not
detectable.
Based
on
the
range
of
MRLs,
0.5
 
1
µ
g/
L
(
ppb),
the
exposure
to
bromoform
is
estimated
to
range
from
1
to
2
µ
g/
day
for
surface
water
supplies.

Ingestion
exposure
to
brominated
trihalomethanes
in
drinking
water
can
also
be
estimated
from
the
concentrations
found
at
the
tap
in
the
U.
S.
EPA's
Total
Exposure
Assessment
Methodology
(
TEAM)
study.
Estimates
of
the
average
of
the
population
intakes
for
ingestion
of
bromodichloromethane
from
drinking
water
range
from
0.42
to
42

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
2.0
to
90

g/
person/
day.
Estimates
of
the
average
population
intake
of
dibromochloromethane
from
drinking
water
range
from
0.2
to
56

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
0.9
to
86

g/
person/
day.
Estimates
of
the
average
of
the
population
intakes
of
bromoform,
for
those
areas
in
which
bromoform
was
measurable
in
a
majority
of
the
samples,
range
from
1.6
to
16.2

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
2.4
to
26

g/
person/
day.
Four
of
the
six
locations
in
the
TEAM
study,
however,
had
a
low
frequency
(
less
than
10%)
of
detection
of
bromoform
in
measurable
quantities.

Sources
of
uncertainty
in
these
estimates
of
ingestion
exposure
include
use
of
different
analytical
methods,
failure
to
report
quantitation
limits,
using
measures
near
the
detection
limit,
failure
to
report
how
nondetects
are
handled
when
averaging
values
(
e.
g.,
set
to
zero
or
one
half
the
detection
limit),
and
failure
to
report
sample
storage
method
and
duration.
In
addition,
many
environmental
factors
influence
the
concentrations
of
these
compounds
in
drinking
water
at
the
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
5
tap
and
in
vended
or
bottled
waters
used
for
drinking.
These
factors
include
season
and
temperature,
geographic
location,
source
of
water,
residence
time
in
distribution
system,
and
others.

Relatively
few
studies
have
analyzed
non­
beverage
foods
for
the
occurrence
of
brominated
trihalomethanes.
In
the
few
studies
available,
bromodichloromethane
has
been
detected
in
nonbeverage
foods
(
i.
e.,
in
one
sample
of
butter
at
7
ppb,
in
three
samples
of
ice­
cream
at
0.6
to
2.3
ppb,
in
6
of
10
samples
of
bean
curd
at
1.2
to
5.2
ppb,
and
in
one
sample
of
bacon).
In
addition,
bromodichloromethane
was
detected
in
one
sample
each
of
eleven
foods
out
of
70
tested
in
14
Market
Baskets
for
the
FDA
Total
Diet
Study.
The
detected
concentrations
ranged
from
10
to
37
ppb
for
individual
food
items.
Studies
that
analyzed
non­
beverage
foods
for
dibromochloromethane
and
bromoform
detected
neither
compound
in
any
of
the
tested
samples.
Brominated
trihalomethanes
have
been
detected
in
up
to
a
third
or
one
half
of
the
types
of
prepared
beverages
examined
in
some
studies,
being
detected
most
frequently
in
colas
and
other
carbonated
soft
drinks.
Bromodichloromethane
has
been
found
most
frequently
of
the
three
compounds
and
bromoform
the
least
frequently.
Bromodichloromethane
was
detected
in
approximately
half
of
the
prepared
beverages
examined
by
McNeal
et
al.
(
1995)
in
the
United
States
and
in
all
of
13
soft
drinks
that
they
analyzed.
One
sampled
soft
drink
contained
bromodichloromethane
at
a
concentration
of
12
ppb;
the
remainder
of
the
samples
contained
less
than
4
ppb.
Bromodichloromethane
was
detected
in
one
sample
of
fruit
juice
at
5
ppb.

Some
data
on
the
occurrence
of
brominated
trihalomethanes
in
foods
and
beverages
are
available
from
studies
conducted
in
Italy,
Japan,
and
Finland.
These
studies
were
also
limited
in
scope
to
examination
of
relatively
few
food
or
beverage
items.
Bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
measured
in
foods
and
beverages
in
Italy,
Japan
and
Finland
ranged
from
undetectable
to
40
ppb,
undetectable
to
13.9
ppb,
and
undetectable
to
10.7
ppb,
respectively.
Because
of
possible
differences
in
water
disinfection
or
food
processing
practices,
these
data
may
not
be
representative
of
concentrations
in
foods
and
beverages
produced
in
the
U.
S.

Measured
concentrations
of
brominated
trihalomethanes
in
outdoor
air
are
variable
from
site
to
site.
When
data
from
several
urban/
suburban
and
source­
dominated
sites
in
Texas,
Louisiana,
North
Carolina
and/
or
Arkansas
were
combined,
the
resulting
average
outdoor
air
concentrations
were
110
ppt
(
0.74
µ
g/
m3)
for
bromodichloromethane,
3.8
ppt
(
0.032
µ
g/
m3)
for
dibromochloromethane,
and
3.6
ppt
(
0.037
µ
g/
m3)
for
bromoform.
A
regional
study
conducted
at
several
sites
in
southern
California
found
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
35%,
17%,
and
31%
of
the
samples,
respectively.
The
maximum
concentrations
observed
were
40
ppt
(
0.27
µ
g/
m3)
for
bromodichloromethane;
290
ppt
(
2.5
µ
g/
m3)
for
dibromochloromethane;
310
ppt
(
3.2
µ
g/
m3)
for
bromoform.
Bromodichloromethane
was
detected
in
64%
(
n=
11)
and
17%
(
n=
6)
of
personal
air
samples
collected
in
Texas
and
North
Carolina.
The
detected
concentrations
ranged
from
0.12
to
4.36
µ
g/
m3
(
0.017
to
0.65
ppb).
Dibromochloromethane
was
not
detected.

Mean
concentrations
in
indoor
air
ranged
from
0.38
to
0.75
µ
g/
m3
for
bromodichloro­
Draft
­
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Cite
or
Quote
February
20,
2003
I
­
6
methane;
0.44
to
0.53
µ
g/
m3
for
dibromochloromethane,
and
0.29
to
0.35
µ
g/
m3
for
bromoform,
as
determined
from
15
minute
samples
collected
in
48
New
Jersey
residences.
It
was
not
clear
whether
these
values
were
based
exclusively
on
detected
concentrations.
In
a
separate
study,
levels
of
brominated
trihalomethanes
in
indoor
air
were
locally
increased
(
e.
g.,
in
shower/
bath
enclosures
and
vanity
areas)
during
showering
and
bathing
events.
The
levels
of
individual
brominated
trihalomethanes
in
air
were
reported
to
be
consistent
with
the
levels
in
tap
water.

No
data
for
occurrence
of
brominated
trihalomethanes
in
soil
were
available
in
the
materials
reviewed
for
this
document.
The
chemical
and
physical
properties
of
the
brominated
trihalomethanes
indicate
that
they
should
volatilize
readily
from
wet
or
dry
soil
surfaces.
Therefore,
ingestion
of
soil
is
not
expected
to
be
a
significant
route
of
exposure.

Brominated
trihalomethanes
have
been
detected
in
the
blood
and
breast
milk
of
humans.
The
level
of
individual
brominated
trihalomethanes
in
blood
increases
shortly
after
exposure
to
these
compounds
in
tap
water
(
by
dermal
contact
and/
or
inhalation
of
the
volatilized
compound)
during
bathing
and
showering.
Dibromochloromethane
was
detected
in
one
of
eight
samples
of
breast
milk
collected
from
women
living
in
the
vicinity
of
U.
S.
chemical
manufacturing
plants
or
user
facilities.

The
RSC
(
relative
source
contribution)
is
the
percentage
of
total
daily
exposure
that
is
attributable
to
tap
water
when
all
potential
sources
are
considered
(
e.
g.,
air,
food,
soil,
and
water).
Ideally,
the
RSC
is
determined
quantitatively
using
nationwide,
central
tendency
and/
or
high­
end
estimates
of
exposure
from
each
relevant
medium.
In
the
absence
of
such
data,
a
default
RSC
ranging
from
20%
to
80%
may
be
used.

The
RSC
used
in
the
current
and
previous
drinking
water
regulations
for
dibromochloromethane
is
80%.
This
value
was
established
by
use
of
a
screening
level
approach
to
estimate
and
compare
exposure
to
dibromochloromethane
from
various
sources.
Information
considered
for
during
this
process
is
summarized
in
Appendix
C.
There
are
some
uncertainties
in
the
80%
RSC
that
are
related
to
the
availability
of
adequate
concentration
data
for
dibromochloromethane
in
media
other
than
water.
Parallel
RSC
calculations
were
not
performed
for
bromodichloromethane
and
bromoform.
The
EPA
has
set
the
regulatory
level
for
these
chemicals
in
drinking
water
at
zero
because
it
has
been
determined
that
they
are
probable
human
carcinogens.
Therefore,
determination
of
an
RSC
is
not
relevant
for
these
chemicals
because
it
is
the
Agency's
policy
to
perform
RSC
analysis
only
for
noncarcinogens.

The
use
of
chlorine
to
disinfect
swimming
pools
and
hot
tubs
results
in
the
formation
of
brominated
trihalomethanes.
Swimming
pool
and
hot
tub
users
are
potentially
exposed
to
brominated
trihalomethanes
via
dermal
contact,
ingestion,
and
inhalation
of
compounds
released
to
the
overlying
air.
As
a
result,
swimming
pool
and
hot
tub
users
may
experience
greater
overall
exposures
to
brominated
trihalomethanes
than
the
general
population.
One
study
indicated
that
bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
in
swimming
pool
and
hot
tub
water
ranged
from
1
to
105
µ
g/
L
(
ppb),
from
0.1
to
48
µ
g/
L
(
ppb),
and
from
less
than
0.1
to
62
µ
g/
L
(
ppb),
respectively.
Concentrations
of
the
same
brominated
trihalomethanes
Draft
­
Do
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Cite
or
Quote
February
20,
2003
I
­
7
in
the
air
two
meters
above
the
pool
water
ranged
from
less
than
0.1
to
14
µ
g/
m3
(
0.015
 
2.09
ppb),
from
less
than
0.1
 
10
µ
g/
m3
(
0.011
 
1.2
ppb),
and
from
less
than
0.1
to
5.0
µ
g/
m3
(
0.0097
 
0.48
ppb),
respectively.
Data
from
several
studies
confirm
the
uptake
of
brominated
trihalomethanes
from
swimming
pools,
hot
tubs,
and
environs
by
dermal
and/
or
inhalation
pathways.

Health
Effects
of
Acute
and
Short­
term
Exposure
of
Animals
Large
oral
doses
of
brominated
trihalomethanes
are
lethal
to
mice
and
rats.
Reported
acute
LD
50
values
range
from
450
to
969
mg/
kg
for
bromodichloromethane,
800
to
1,200
mg/
kg
for
dibromochloromethane,
and
1,388
to
1,550
mg/
kg
for
bromoform.

Acute
oral
exposure
to
sublethal
doses
of
brominated
trihalomethanes
can
also
produce
effects
on
the
central
nervous
system,
liver,
kidney,
and
heart.
Ataxia,
anaesthesia,
and/
or
sedation
were
noted
in
mice
receiving
500
mg/
kg
bromodichloromethane,
500
mg/
kg
dibromochloromethane,
or
1,000
mg/
kg
bromoform.
Renal
tubule
degeneration,
necrosis,
and
elevated
levels
of
urinary
markers
of
renal
toxicity
have
been
observed
in
rats
receiving
200
to
400
mg/
kg
bromodichloromethane.
Elevated
levels
of
serum
markers
for
hepatotoxicity
and
have
been
observed
in
rats
at
doses
of
bromodichloromethane
ranging
from
approximately
82
to
400
mg/
kg­
day,
and
hepatocellular
degeneration
and
necrosis
were
observed
at
400
mg/
kg.
Effects
on
heart
contractility
were
reported
in
male
rats
at
doses
of
333
and
667
mg/
kg
dibromochloromethane.

Short­
term
oral
exposure
of
laboratory
animals
to
brominated
trihalomethanes
has
been
observed
to
cause
effects
on
the
liver
and
kidney.
Hepatic
effects,
including
organ
weight
changes,
elevated
serum
enzyme
levels,
and
histopathological
changes,
became
evident
in
mice
and/
or
rats
administered
38
to
250
mg/
kg­
day
bromodichloromethane,
147
to
500
mg/
kg­
day
dibromochloromethane,
or
187
to
289
mg/
kg­
day
bromoform
for
14
to
30
days.
Kidney
effects,
characterized
by
decreased
p­
aminohippurate
uptake,
histopathological
changes,
and
organ
weight
changes,
became
evident
in
mice
and/
or
rats
administered
148
to
300
mg/
kg­
day
bromodichloromethane,
147
to
500
mg/
kg­
day
dibromochloromethane,
or
289
mg/
kg­
day
bromoform
for
14
days.
Evidence
for
decreased
immune
function
was
noted
at
bromodichloromethane
and
dibromochloromethane
doses
of
125
mg/
kg­
day.

The
inhalation
toxicity
of
bromodichloromethane
has
been
evaluated
in
wild
type
and
p53
heterozygous
FVB/
N
and
C57BL/
N
mice.
Dose­
related
renal
tubular
degeneration,
and
associated
regenerative
cell
proliferation
were
seen
in
all
strains
at
concentrations
of
10
ppm
and
above
after
one
week
of
exposure.
Dose­
related
increases
in
hepatic
degeneration
and
regenerative
cell
proliferation
were
observed
at
30
ppm
and
above.
After
three
weeks
of
exposure,
macroscopic
and
histologic
lesions
in
the
kidney
and
liver
were
resolved
and
cell
proliferation
indices
had
returned
to
near
baseline
levels.
Pathological
changes
were
more
severe
in
the
FVB/
N
compared
to
the
C57BL/
N
mice
and
were
more
severe
in
the
heterozygotes
than
in
the
wild
type
strains.
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I
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8
Health
Effects
of
Subchronic
and
Chronic
Exposure
of
Animals
The
predominant
effects
of
subchronic
oral
exposure
occur
in
the
liver
and
kidney.
The
effects
produced
in
these
two
organs
are
similar
in
nature
to
those
described
for
short­
term
exposures,
with
liver
appearing
to
be
the
most
sensitive
target
organ
for
dibromochloromethane
and
bromoform
exposure.
Histopathological
changes
in
the
liver
were
reported
in
mice
and/
or
rats
administered
200
mg/
kg­
day
bromodichloromethane,
50
to
250
mg/
kg­
day
dibromochloromethane
or
50
to
283
mg/
kg­
day
bromoform.
Histopathological
changes
in
the
kidney
were
reported
in
mice
and/
or
rats
administered
100
mg/
kg­
day
bromodichloromethane,
or
250
mg/
kgday
dibromochloromethane.

As
observed
for
shorter
durations
of
exposure,
the
predominant
effects
of
chronic
oral
exposure
are
observed
in
the
liver
and
kidney.
Histopathological
signs
of
hepatic
toxicity
in
mice
and/
or
rats
became
evident
at
doses
of
6
to
50
mg/
kg­
day
for
bromodichloromethane,
40
to
50
mg/
kg­
day
for
dibromochloromethane,
and
90
to
152
mg/
kg­
day
for
bromoform.
Signs
of
bromodichloromethane­
induced
renal
toxicity
became
evident
in
mice
and
rats
treated
with
doses
of
25
and
50
mg/
kg­
day,
respectively.

Reproductive/
Developmental
Effects
in
Animals
Reproductive
and
developmental
studies
of
brominated
trihalomethanes
are
summarized
in
Table
V­
9.
Data
on
the
developmental
effects
of
brominated
trihalomethanes
suggest
that
these
chemicals
are
toxic
to
the
fetus
in
most
cases
only
at
doses
that
result
in
maternal
toxicity.
Signs
of
maternal
toxicity
(
decreased
body
weight,
body
weight
gain
and/
or
changes
in
organ
weight)
were
reported
in
rats
administered
bromodichloromethane
at
25
to
200
mg/
kg­
day
and
in
rabbits
administered
4.9
to
35.6
mg/
kg­
day.
Signs
of
maternal
toxicity
were
observed
in
rats
or
mice
administered
17
(
marginal)
to
200
mg/
kg­
day
dibromochloromethane
and
in
mice
administered
100
mg/
kg­
day
bromoform.
Maternal
toxicity
was
not
observed
in
female
rats
dosed
with
up
to
200
mg/
kg­
day
of
bromoform.
Several
well­
conducted
studies
on
the
developmental
toxicity
of
bromodichloromethane
gave
negative
results
at
doses
up
to
116
mg/
kg­
day
in
rats
and
76
mg/
kgday
in
rabbits
when
administered
in
drinking
water.
However,
in
other
studies,
slightly
decreased
numbers
of
ossification
sites
in
the
hindlimb
and
forelimb
were
observed
in
fetuses
of
Sprague­
Dawley
rats
administered
45
mg/
kg­
day
in
the
drinking
water
on
gestation
days
6
to
21and
sternebral
aberrations
were
observed
in
the
offspring
of
Sprague­
Dawley
rats
administered
200
mg/
kg­
day
by
gavage
in
corn
oil.
Reductions
in
mean
pup
weight
gain
and
pup
weight
were
observed
when
the
pups
were
administered
bromodichloromethane
in
the
drinking
water
at
concentrations
of
150
ppm
and
above
(
biologically
meaningful
estimates
of
intake
on
a
mg/
kg­
day
basis
could
not
be
calculated
for
this
study).
Full
litter
resorption
has
been
noted
in
F344
rats,
but
not
Sprague­
Dawley
rats,
treated
with
bromodichloromethane
at
doses
of
50
to
100
mg/
kg­
day
during
gestation
days
6
to
10.
Chronic
oral
exposure
of
male
F344
rats
to
bromodichloromethane
resulted
in
reduced
sperm
velocities
at
doses
of
39
mg/
kg­
day.
This
response
was
not
accompanied
by
histopathological
changes
in
any
reproductive
tissue
examined.
Adverse
clinical
signs
and
reduced
body
weight
and
body
weight
gain
were
observed
in
parental
generation
female
rats
and
F
1
male
and
female
rats
at
150
ppm
(
approximately
11.6
to
40.2
mg/
kg­
day)
in
a
two
Draft
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2003
I
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9
generation
drinking
water
study
of
bromodichloromethane.
In
the
same
study,
small
but
statistically
significant
delays
in
sexual
maturation
occurred
in
F1
males
at
50
ppm
(
approximately
11.6
to
40.2
mg/
kg­
day)
and
F1
females
at
450
ppm
(
approximately
29.5
to
109
mg/
kg­
day).
These
delays
may
have
been
secondary
to
dehydration
caused
by
taste
aversion
to
bromodichloromethane
in
the
drinking
water.

Four
of
five
studies
on
the
reproductive
or
developmental
toxicity
of
dibromochloromethane
gave
negative
results
when
tested
at
doses
of
up
to
200
mg/
kg­
day.
In
the
fifth
study,
dibromochloromethane
administered
at
17
mg/
kg­
day
in
a
multigenerational
study
resulted
in
reduced
day
14
postnatal
body
weight
in
one
of
two
F2
generation
litters.
At
171
mg/
kg­
day,
the
mid­
dose
in
the
study,
decreased
litter
size,
viability
index,
lactation
index,
and
postnatal
body
weight
were
observed
in
some
F1
and/
or
F2
generations.

The
developmental
and
reproductive
toxicity
of
bromoform
has
been
examined
in
two
studies.
Bromoform
administered
to
Sprague­
Dawley
rats
at
100
mg/
kg­
day
in
corn
oil
by
gavage
resulted
in
a
significant
increase
in
sternebral
aberrations
in
the
apparent
absence
of
maternal
toxicity.
In
a
continuous
breeding
toxicity
protocol,
gavage
doses
of
200
mg/
kg­
day
in
corn
oil
resulted
decreased
postnatal
survival,
organ
weight
changes,
and
liver
histopathology
in
F1
ICR
Swiss
mice
of
both
sexes.
No
effects
on
fertility
or
other
reproductive
endpoints
were
noted.

Mutagenicity
and
Carcinogenicity
Studies
In
vitro
and
in
vivo
studies
of
the
mutagenic
and
genotoxic
potential
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
have
yielded
mixed
results.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid),
and
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
Overall,
a
majority
of
studies
yielded
more
positive
results
for
bromoform
and
bromodichloromethane.
The
genotoxicity
and
mutagenicity
data
for
dibromochloromethane
are
inconclusive.
Recent
studies
in
strains
of
Salmonella
engineered
to
contain
rat
theta­
class
glutathione
S­
transferase
suggest
that
mutagenicity
of
the
brominated
trihalomethanes
may
be
mediated
by
glutathione
conjugation.

Carcinogenicity
Studies
in
Animals
The
carcinogenic
potential
of
individual
brominated
trihalomethanes
has
been
investigated
in
oral
exposure
studies
conducted
in
mice
and
rats.
Brominated
trihalomethanes
administered
individually
in
drinking
water
induced
the
formation
of
aberrant
crypt
foci
(
ACF),
putative
preneoplastic
lesions,
in
the
colons
of
male
F344/
N
rats,
but
not
male
B6C3F
1
mice.
However,
a
cancer
bioassay
of
bromodichloromethane
administered
in
the
drinking
water
did
not
detect
colon
cancer
in
male
F344/
N
rats.
Ingestion
of
bromodichloromethane
in
corn
oil
significantly
increased
the
incidence
of
liver
tumors
in
female
mice,
renal
tumors
in
male
mice
and
in
male
and
female
rats,
and
tumors
of
the
large
intestine
in
male
and
female
rats.
A
drinking
water
exposure
study
of
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20,
2003
I
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10
bromodichloromethane
in
rats
reported
a
significant
induction
of
liver
tumors
at
the
lowest
dose
tested,
but
not
at
higher
doses.
Administration
of
dibromochloromethane
in
corn
oil
significantly
increased
the
incidence
of
liver
tumors
in
male
and
female
mice.
Administration
of
bromoform
in
corn
oil
increased
the
incidence
of
uncommon
intestinal
tumors
in
male
and
female
rats,
with
the
effect
reaching
statistical
significance
in
females.
No
data
are
available
for
the
carcinogenicity
of
brominated
trihalomethanes
via
the
inhalation
or
dermal
routes.

Other
Key
Health
Effects
Data
from
Animal
Studies
The
immunotoxicity
of
brominated
trihalomethanes
has
been
investigated
in
mice
and
rats.
Short­
term
bromodichloromethane
exposure
resulted
in
decreased
antibody
forming
cells
in
serum,
decreased
hemagglutinin
titers,
and/
or
suppression
of
Con
A­
stimulated
proliferation
of
spleen
cells
at
doses
of
125
to
250
mg/
kg­
day.

Studies
in
pregnant
F344
rats
detected
decreased
levels
of
progesterone
in
animals
administered
75
or
100
mg/
kg
bromodichloromethane
by
aqueous
gavage
on
gestation
day
8
or
9.
Increased
levels
of
luteinizing
hormone
were
observed
two
to
three
days
after
dose
administration.
Disruption
of
luteal
responsiveness
to
luteinizing
hormone
has
been
proposed
as
a
possible
mode
of
action
by
which
bromodichloromethane
elicits
full
litter
resorption
in
F344
rats,
but
additional
data
are
required
to
confirm
this
hypothesis.
The
available
evidence
does
not
suggest
that
dibromochloromethane
or
bromoform
affect
hormonal
parameters.

Limited
structure­
activity
data
for
brominated
trihalomethanes
and
chloroform
suggest
that
bromination
may
influence
the
proportion
of
compound
metabolized
via
the
oxidative
and
reductive
pathways,
with
brominated
compounds
being
more
extensively
metabolized
by
the
reductive
pathway.
Additional
evidence
suggests
that
a
GST­
mediated
pathway
may
play
an
important
role
in
metabolism
of
brominated
trihalomethanes.

Health
Effects
in
Humans
Limited
human
health
data
are
available
for
the
brominated
trihalomethanes.
In
the
past,
bromoform
was
used
as
a
sedative
for
children
with
whooping
cough.
Doses
of
50
to
100
mg/
kgday
usually
produced
sedation
without
apparent
adverse
effects.
Some
rare
instances
of
death
or
near­
death
were
reported,
although
these
cases
were
generally
attributed
to
accidental
overdoses.
No
human
toxicological
data
were
available
for
bromodichloromethane
or
dibromochloromethane.

Numerous
epidemiological
studies
have
examined
the
association
between
water
chlorination
and
increased
cancer
mortality
rates.
None
of
these
studies
has
identified
a
strong
association
between
cancer
and
exposure
to
any
individual
brominated
trihalomethane.

Several
epidemiological
studies
have
examined
the
association
of
chlorinated
water
use
with
various
pregnancy
outcomes,
including
low
birth
weight,
premature
birth,
intrauterine
growth
retardation,
spontaneous
abortion,
stillbirth
,
and
birth
defects.
An
association
has
been
Draft
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February
20,
2003
I
­
11
reported
for
exposure
to
bromodichloromethane
(
or
a
closely
associated
compound)
and
a
moderately
increased
risk
of
spontaneous
abortion
during
the
first
trimester.
A
confirmation
of
this
finding
is
pending
reanalysis
of
the
original
data
to
correct
a
differential
misclassification
error
identified
in
a
subsequent
analysis
of
the
study
data.
An
association
has
also
been
reported
for
exposure
to
bromodichloromethane
(
or
a
closely
associated
compound)
and
1)
stillbirth
of
fetuses
weighing
more
than
500
g
and
2)
increased
risk
of
neural
tube
defects
in
women
exposed
to

20
µ
g/
L
of
bromodichloromethane
prior
to
conception
through
the
first
month
of
pregnancy.
An
association
has
been
reported
for
total
brominated
trihalomethanes
and
reduced
menstrual
cycle
and
follicular
phase
length
in
women
of
child­
bearing
age.
Among
the
individual
brominated
trihalomethanes,
dibromochloromethane
displayed
the
strongest
association
with
altered
menstrual
function.
However,
because
chlorinated
water
contains
many
disinfection
by­
products,
it
is
not
possible
to
directly
conclude
from
these
studies
that
bromodichloromethane
or
bromodichloromethane
are
developmental
toxicants
in
humans.
Nevertheless,
these
studies
raise
significant
concern
for
possible
human
health
effects.
The
methodology
used
to
estimate
exposure
to
brominated
trihalomethanes
in
tap
water
has
been
examined
with
the
goal
of
refining
estimates
of
intake
of
these
compounds
in
epidemiological
studies.

Susceptible
Populations
There
is
currently
no
clear
evidence
to
suggest
that
children
or
the
fetus
are
at
greater
risk
for
adverse
effects
from
exposure
to
bromoform
or
dibromochloromethane
than
are
adults.
Where
evidence
of
developmental
toxicity
has
been
observed
in
animals
exposed
to
these
chemicals,
it
generally
appears
to
occur
only
at
doses
that
elicit
signs
of
maternal
toxicity.
Associations
between
concentration
of
bromodichloromethane
(
or
a
co­
occurring
chemical)
and
spontaneous
abortion
or
still
birth
have
been
observed
in
two
epidemiological
studies.
Evidence
in
rats
indicates
that
exposure
to
bromodichloromethane
causes
full
litter
resorption
in
F344
rats
but
not
Sprague­
Dawley
rats,
at
doses
many­
fold
higher
than
are
likely
to
occur
from
human
ingestion
of
drinking
water.
Full
litter
absorption
appears
to
result
from
a
maternally­
mediated
mode
of
action,
rather
than
from
a
direct
effect
on
the
embryo.
A
mechanism
of
action
for
bromodichloromethane­
related
pregnancy
loss
is
suggested
for
the
rat
(
reduced
sensitivity
of
the
corpus
luteum
to
luteinizing
hormone),
but
is
not
without
alternative
explanation.
At
present,
there
is
insufficient
information
on
the
mode
of
action
leading
to
full
litter
resorption
in
rats
to
fully
evaluate
the
relevance
of
this
outcome
to
potential
reproductive
and/
or
developmental
toxicity
in
humans.
There
is
presently
no
evidence
to
suggest
that
or
the
fetus
are
at
increased
risk
for
brominated
trihalomethane
toxicity
as
a
result
of
higher
levels
of
metabolizing
enzymes.

Subpopulations
with
either
high
levels
of
glutathione
S­
transferase
or
low
baseline
levels
of
glutathione
may
potentially
be
more
sensitive
than
the
general
population
to
brominated
trihalomethane­
induced
toxicity,
but
there
are
currently
no
epidemiological
or
animal
data
to
confirm
this
speculation.
The
functional
significance
of
polymorphisms
in
cytochrome
P450
isoforms
that
metabolize
brominated
trihalomethanes
is
also
unknown.
Populations
that
have
higher
levels
of
CYP2E1
activity
as
a
result
of
co­
exposure
to
ethanol
or
other
inducers
or
as
a
result
of
altered
health
or
physiological
states
may
experience
potentially
higher
risk.
Draft
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20,
2003
I
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12
Mechanism
of
Toxicity
It
is
generally
believed
that
the
toxicity
of
the
brominated
trihalomethanes
is
related
to
their
metabolism.
This
conclusion
is
based
largely
on
the
observation
that
liver
and
kidney,
the
chief
target
tissues
for
these
compounds,
are
also
the
primary
sites
of
their
metabolism.
In
addition,
treatments
which
increase
or
decrease
metabolism
also
tend
to
increase
or
decrease
trihalomethane­
induced
toxicity
in
parallel.

Metabolism
of
brominated
trihalomethanes
is
believed
to
occur
via
oxidative
and
reductive
pathways.
Limited
structure­
activity
data
for
brominated
trihalomethanes
and
the
structurallyrelated
trihalomethane
chloroform
suggest
that
bromination
may
influence
the
proportion
of
compound
metabolized
via
the
oxidative
and
reductive
pathways,
with
brominated
compounds
being
more
extensively
metabolized
by
the
reductive
pathway.
Additional
evidence
suggests
that
a
GSH­
mediated
pathway
may
play
an
important
role
in
metabolism
of
brominated
trihalomethanes.
These
data
raise
the
possibility
that
brominated
trihalomethanes
may
induce
adverse
effects
(
toxicity
and
carcinogenicity)
via
several
different
pathways.

The
precise
biochemical
mechanisms
which
link
brominated
trihalomethane
metabolism
to
toxicity
have
not
been
characterized,
but
many
researchers
have
proposed
that
toxicity
results
from
the
production
of
reactive
intermediates.
Reactive
intermediates
may
arise
from
either
the
oxidative
(
dihalocarbonyls)
or
the
reductive
(
free
radicals)
pathways
of
metabolism.
Such
reactive
intermediates
are
known
to
form
covalent
adducts
with
various
cellular
molecules,
and
may
impair
the
function
of
those
molecules
and
cause
cell
injury.
Free
radical
production
may
also
lead
to
cell
injury
by
inducing
lipid
peroxidation
in
cellular
membranes.
Direct
evidence
showing
a
relationship
between
the
level
of
covalent
binding
intermediates
generated
by
either
pathway
and
the
extent
of
toxicity
is
not
available
for
the
brominated
trihalomethanes.
Manipulation
of
cellular
glutathione
levels
suggests
that
this
compound
may
play
a
protective
role
in
brominated
trihalomethane­
induced
toxicity.

Individual
brominated
trihalomethanes
have
been
shown
to
induce
tumors
in
laboratory
animals.
The
mode
of
action
by
which
brominated
trihalomethanes
induce
tumors
in
target
tissues
has
not
been
fully
characterized.
DNA
adducts
can
be
formed
by
interaction
of
reactive
metabolites
(
resulting
from
oxidative
and
reductive
metabolism)
with
DNA.
In
addition,
preliminary
evidence
suggests
that
DNA
adducts
can
be
formed
through
conjugation
with
glutathione
and
bioactivation
of
the
resulting
conjugates.
In
contrast
to
chloroform,
the
role
of
cytotoxicity
and
associated
regenerative
cell
proliferation
in
tumorigenicity
of
brominated
trihalomethanes
is
presently
unclear.
Comparison
of
dose­
response
data
for
liver
toxicity
(
including
cell
proliferation)
and
tumorigenicity
in
mice
suggests
that
tumor
formation
occurs
at
concentrations
lower
than
those
which
stimulate
cell
proliferation.
No
evidence
for
increased
cell
proliferation
in
kidney
was
obtained
in
studies
using
doses
up
to
246
mg/
kg­
day
for
bromodichloromethane,
312
mg/
kg­
day
for
dibromochloromethane,
or
379
mg/
kg­
day
for
bromoform.

Interaction
with
agents
which
increase
or
decrease
the
activity
of
enzymes
responsible
for
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
13
metabolism
of
brominated
trihalomethanes
may
modify
carcinogenicity/
toxicity.
Pretreatment
with
inducers
of
CYP2E1
has
been
observed
to
increase
the
hepatotoxicity
of
bromodichloromethane
and
dibromochloromethane
in
male
rats.
Pretreatment
with
m­
xylene,
an
inducer
of
the
CYP2B1/
CYP2B2
isoforms,
increased
the
hepatotoxicity
of
dibromochloromethane
in
male
rats.
Conversely,
administration
of
the
cytochrome
P450
inhibitor
1­
aminobenzotriazole
prevented
bromodichloromethane­
induced
hepatotoxicity
in
rats.
Recent
findings
indicating
possible
glutathione­
mediated
metabolism
of
brominated
trihalomethanes
suggest
that
treatments
or
agents
which
alter
glutathione­
S­
transferase
activity
could
potentially
modify
the
toxicity
of
brominated
trihalomethanes.

The
severity
of
brominated
trihalomethane
toxicity
is
potentially
affected
by
the
vehicle
of
administration.
In
a
study
of
vehicle
effects
on
the
acute
toxicity
of
bromodichloromethane,
a
high
dose
(
400
mg/
kg)
of
the
chemical
was
more
hepato­
and
nephrotoxic
when
given
in
corn
oil
compared
to
aqueous
administration,
but
this
difference
was
not
evident
at
a
lower
dose
(
200
mg/
kg).

Quantification
of
Noncarcinogenic
Effects
Candidate
health
effects
endpoints
were
analyzed
by
benchmark
dose
(
BMD)
modeling
using
a
benchmark
response
of
10%
added
risk.
The
BMDL
10
was
defined
as
the
95%
lower
bond
on
the
BMD
estimate.
For
bromodichloromethane,
a
BMDL
10
of
30
mg/
kg­
day
identified
on
the
basis
of
full
litter
resorption
in
F344
rats
was
used
to
calculate
a
One­
day
Health
Advisory
(
HA)
value
of
1
mg/
L.
A
BMDL
10
of
18
mg/
kg­
day
for
single
cell
hepatic
necrosis,
identified
in
a
30­
day
drinking
water
study
in
rats,
was
used
to
calculate
a
Ten­
day
HA
value
of
0.6
mg/
L.
A
BMDL
10
of
18
mg/
kg­
day
for
reduced
maternal
body
weight
gain
on
gestation
days
6­
9,
identified
in
a
developmental
study
in
rats,
was
used
to
calculate
a
Longer­
term
HA
of
0.6
mg/
L
for
a
10­
kg
child.
A
Longer­
term
HA
value
of
2
mg/
L
was
calculated
for
a
70­
kg
adult
based
on
the
same
endpoint.
The
calculations
for
the
Reference
Dose
(
RfD)
of
0.003
mg/
kg­
day
and
Drinking
Water
Exposure
Level
(
DWEL)
of
90
µ
g/
L
employed
a
duration
adjusted
BMDL
10
of
0.8
mg/
kgday
for
fatty
degeneration
of
the
liver,
identified
in
a
24
month
dietary
study
in
rats.
Because
bromodichloromethane
is
classified
as
a
probable
human
carcinogen,
a
Lifetime
HA
is
not
recommended.

For
dibromochloromethane,
no
suitable
study
was
located
for
the
calculation
of
a
One­
day
HA
value.
Use
of
the
10­
day
HA
value
as
a
conservative
estimate
is
recommended.
The
Ten­
day
HA
value
of
0.6
mg/
L
was
calculated
using
a
BMDL
10
of
5.5
mg/
kg­
day
for
hepatic
cell
vacuolization,
identified
in
a
28­
day
feeding
study
in
rats.
A
duration­
adjusted
BMDL
10
value
of
1.7
mg/
kg­
day
for
hepatic
cell
vacuolization,
identified
in
a
13­
week
gavage
study
in
rats,
was
used
to
calculate
Longer­
term
HA
values
of
0.2
and
0.6
mg/
L
for
a
10­
kg
child
and
a
70­
kg
adult,
respectively.
A
duration­
adjusted
BMDL
10
value
of
1.6
mg/
kg­
day
for
fatty
changes
identified
in
a
2
year
gavage
study
in
mice
was
used
to
calculate
a
RfD
of
0.02
mg/
kg­
day
and
a
DWEL
of
700
µ
g/
L.
The
Lifetime
HA
for
dibromochloromethane
is
60
µ
g/
L.
This
value
was
calculated
using
the
default
RSC
value
of
80%
for
exposure
via
ingestion
of
drinking
water.
Because
this
compound
is
classified
as
a
possible
human
carcinogen,
the
derivation
of
the
Lifetime
HA
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
14
incorporated
an
uncertainty
factor
of
10.

For
bromoform,
an
estimated
dose
of
54
mg/
kg­
day
that
caused
sedation
in
children
was
used
to
calculate
a
One­
day
HA
value
of
5
mg/
L.
A
BMDL
10
of
2.3
mg/
kg­
day
for
hepatic
vacuolization,
identified
in
a
one
month
dietary
study
in
rats,
was
used
to
calculate
a
value
of
0.2
mg/
L
for
the
Ten­
day
HA
for
the
10­
kg
child.
This
value
was
also
recommended
for
use
as
the
Longer­
term
HA
for
a
10
kg
child.
A
duration­
adjusted
BMDL
10
value
of
2.6
mg/
kg­
day
for
hepatic
vacuolization,
identified
in
a
13
week
gavage
study
in
rats,
was
also
used
to
calculate
a
value
of
0.9
mg/
L
for
the
Longer­
term
HA
for
the
70­
kg
adult.
The
BMDL
10
value
of
2.6
mg/
kgday
was
also
used
to
calculate
an
RfD
of
0.03
mg/
kg­
day
and
a
DWEL
of
1000
µ
g/
L.
Because
bromoform
is
classified
as
a
probable
human
carcinogen,
a
Lifetime
HA
is
not
recommended.

Quantification
of
Carcinogenic
Effects
Chronic
oral
exposure
studies
performed
by
the
National
Toxicology
Program
in
rats
and
mice
provide
adequate
data
to
derive
quantitative
cancer
risk
estimates
for
the
three
brominated
trihalomethanes,
although
the
chemicals
were
administered
in
a
corn
oil
vehicle.
For
bromodichloromethane
a
unit
risk
of
1.0
x
10­
6
(
µ
g/
L)­
1
was
derived,
based
on
the
incidence
of
renal
tumors
in
male
mice.
The
oral
slope
factor
and
concentration
for
excess
cancer
risk
of
10­
6
were
3.5
x
10­
2
(
mg/
kg­
day)­
1
and
1.0
µ
g/
L,
respectively.
For
dibromochloromethane,
a
unit
risk
of
1.2
x
10­
6
(
µ
g/
L)­
1
was
derived,
based
on
liver
tumors
in
female
mice.
The
oral
slope
factor
and
concentration
for
excess
cancer
risk
of
10­
6
were
4.3
×
10­
2
(
mg/
kg­
day)­
1
and
0.8
µ
g/
L,
respectively.
For
bromoform,
a
unit
risk
of
1.3
x
10­
7
(
µ
g/
L)­
1
was
derived,
based
on
tumors
of
the
large
intestine
in
female
rats.
The
oral
slope
factor
and
concentration
for
excess
cancer
risk
of
10­
6
were
4.56
×
10­
3
(
mg/
kg­
day)­
1
and
8
µ
g/
L,
respectively.
These
values
were
calculated
using
an
animal­
to­
human
scaling
factor
of
body
weight3/
4
in
accordance
with
proposed
U.
S.
EPA
guidance
(
U.
S.
EPA,
1996;
1999).

In
a
previous
assessment
of
the
carcinogenicity
of
brominated
trihalomethanes,
the
Carcinogenic
Risk
Assessment
Verification
Endeavor
(
CRAVE)
group
of
the
U.
S.
EPA
assigned
bromodichloromethane
and
bromoform
to
Group
B2:
probable
human
carcinogen.
CRAVE
assigned
dibromochloromethane
to
Group
C:
possible
human
carcinogen.
Under
the
proposed
1999
U.
S.
EPA
Guidelines
for
Cancer
Assessment,
bromodichloromethane
and
bromoform
are
likely
to
be
carcinogenic
to
humans
by
all
routes
of
exposure.
This
descriptor
is
appropriate
when
the
available
tumor
data
and
other
key
data
are
adequate
to
demonstrate
carcinogenic
potential
to
humans.
This
finding
is
based
on
the
weight
of
experimental
evidence
in
animal
models
which
shows
carcinogenicity
by
modes
of
action
that
are
relevant
to
humans.
Dibromochloromethane
shows
suggestive
evidence
of
carcinogenicity,
but
not
sufficient
to
assess
human
carcinogenic
potential.
This
descriptor
is
used
when
the
evidence
from
human
or
animal
data
is
suggestive
of
carcinogenicity,
which
raises
a
concern
for
carcinogenic
effects
but
is
not
judged
sufficient
for
a
conclusion
as
to
human
carcinogenic
potential.
This
finding
is
based
on
the
weight
of
experimental
evidence
in
animal
models
which
indicate
limited
or
equivocal
evidence
of
carcinogenicity.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
15
IARC
has
recently
re­
evaluated
the
carcinogenic
potential
of
the
brominated
trihalomethanes.
IARC
classified
bromodichloromethane
as
a
Group
2B
carcinogen:
possibly
carcinogenic
to
humans.
IARC
classified
dibromochloromethane
and
bromoform
as
Group
3:
not
classifiable
as
to
carcinogenicity
in
humans.

Table
I­
1
summarizes
the
quantification
of
noncarcinogenic
and
carcinogenic
effects
for
brominated
trihalomethanes.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
I
­
16
Table
I­
1
Summary
of
Quantification
of
Toxicological
Effects
for
Brominated
Trihalomethanes
Advisory
Value
Reference
Bromodichloromethane
One­
day
HA
for
10­
kg
child
1
mg/
L
Narotsky
et
al.
(
1997)

Ten­
day
HA
for
10­
kg
child
0.6
mg/
L
NTP
(
1998)

Longer­
term
HA
for
10­
kg
child
0.6
mg/
L
CCC
(
2000d)

Longer­
term
HA
for
70­
kg
adult
2
mg/
L
CCC
(
2000d)

RfD
0.003
mg/
kg­
day
Aida
et
al.
(
1992b)

DWEL
100
µ
g/
L
Aida
et
al.
(
1992b)

Lifetime
HA
Not
applicable
­­

Oral
Slope
Factor
c
3.5
x
10­
2
(
mg/
kg­
day)­
1
NTP
(
1987)

Concentration
for
excess
cancer
risk
(
10­
6)
1.0
µ
g/
L
NTP
(
1987)

Unit
Risk
1x10­
6
(
µ
g/
L)­
1
NTP
(
1987)

Dibromochloromethane
One­
day
HA
for
10­
kg
child
b
0.6
mg/
L
Aida
et
al.
(
1992a)

Ten­
day
HA
for
10­
kg
child
0.6
mg/
L
Aida
et
al.
(
1992a)

Longer­
term
HA
for
10­
kg
child
0.2
mg/
L
NTP
(
1985)

Longer­
term
HA
for
70­
kg
adult
0.6
mg/
L
NTP
(
1985)

RfD
0.02
mg/
kg­
day
NTP
(
1985)

DWEL
700
µ
g/
L
NTP
(
1985)

Lifetime
HA
60
µ
g/
L
NTP
(
1985)

Oral
Slope
Factor
c
4.3
x
10­
2
(
mg/
kg­
day)­
1
NTP
(
1985)

Concentration
for
Excess
cancer
risk
(
10­
6)
0.8
µ
g/
L
NTP
(
1985)

Unit
Risk
1.2
x
10­
6
(
µ
g/
L)­
1
NTP
(
1985)

Bromoform
One­
day
HA
for
10­
kg
child
5
mg/
L
Burton­
Fanning
(
1901)

Ten­
day
HA
for
10­
kg
child
0.2
mg/
L
NTP
(
1989a)

Longer­
term
HA
for
10­
kg
child
a
0.2
mg/
L
NTP
(
1989a)

Longer­
term
HA
for
70­
kg
adult
0.9
mg/
L
NTP
(
1989a)

RfD
0.03
mg/
kg­
day
NTP
(
1989a)
Table
I­
1
(
cont.)

Advisory
Value
Reference
Draft
­
Do
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or
Quote
February
20,
2003
I
­
17
DWEL
300
µ
g/
L
NTP
(
1989a)

Lifetime
HA
Not
applicable
­­

Oral
Slope
Factor
c
4.6
×
10­
3
(
mg/
kg­
day)­
1
NTP
(
1989a)

Concentration
for
Excess
cancer
risk
(
10­
6)
8
µ
g/
L
NTP
(
1989a)

Unit
Risk
1.3
x
10­
7
(
µ
g/
L)­
1
NTP
(
1989a)

a
The
calculated
value
for
the
Longer­
term
HA
was
slightly
higher
than
the
values
for
the
Ten­
day
HAs.
Therefore,
use
of
the
Ten­
day
HA
for
a
10­
kg
child
is
recommended
as
an
estimate
of
the
Longer­
term
HA
for
a
10­
kg
child.
b
Use
of
the
Ten­
day
HA
recommended
as
an
estimate
of
the
One­
day
HA
for
a
10­
kg
child.
c
The
oral
slope
factor
was
calculated
using
the
Linearized
Multistage
model
(
extra
risk)
and
an
animal­
to­
human
scaling
factor
of
body
weight3/
4
Abbreviations:
BW,
Body
weight;
DWEL,
Drinking
water
exposure
limit;
HA,
Health
advisory;
LMS;
Linearized
Multistage
Model
Draft
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February
20,
2003
II
­
1
C
H
Cl
Cl
Br
C
H
Br
Cl
Br
C
H
Br
Br
Br
II.
PHYSICAL
AND
CHEMICAL
PROPERTIES
A.
Properties
and
Uses
Bromodichloromethane
(
CHBrCl
2),
dibromochloromethane
(
CHBr
2
Cl)
and
bromoform
(
CHBr
3)
are
clear
liquids
with
higher
densities
than
the
structurally­
related
compound
chloroform.
They
have
limited
solubility
in
water
but
are
very
soluble
in
organic
solvents
(
Windholz,
1976).
Some
important
physical
and
chemical
properties
of
these
bromine­
containing
trihalomethanes
are
summarized
in
Table
II­
1.
Brominated
trihalomethanes
are
sufficiently
volatile
to
evaporate
from
drinking
water
(
Jolley
et
al.,
1978).

Table
II­
1
Physical
and
Chemical
Properties
of
the
Brominated
Trihalomethanes
Property
Chemical
Bromodichloromethane
Dibromochloromethane
Bromoform
Structure
Chemical
Abstracts
Registry
Number
75­
27­
4
124­
48­
1
75­
25­
2
Registry
of
Toxic
Effects
of
Chemical
Substances
Number
PA
5310000
PA
6360000
PB
5600000
Synonyms
dichlorobromomethane
chlorodibromomethane
tribromomethane
Chemical
Formula
CHBrCl2
CHBr2Cl
CHBr3
Molecular
Weight
163.83
208.29
252.77
Boiling
Point
90
°
C
116C
149
­
150
°
C
Melting
Point
­
57.1
°
C
­­
6­
7
°
C
Specific
Gravity
(
20
°
)
1.980
2.38
2.887
Vapor
Pressure
50mm
(
20
°
C)
15
mm
(
10
°
C)
5.6
mm
(
25
°
C)

Stability
in
Water
volatile
volatile
volatile
Water
Solubility
3,032
mg/
L
(
30
°
C)
1,050
mg/
L
(
30
°
C)
3,190
mg/
L
(
30
°
C)

Log
Octanol:
Water
Partition
Coefficient
(
Kow)
2.09
2.23
2.37
Draft
­
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Not
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or
Quote
February
20,
2003
II
­
2
Source:
USEPA
(
1992c;
1994b)

Brominated
trihalomethanes
also
occur
in
drinking
water
as
by­
products
of
chlorination.
Bromide
(
Br­),
a
common
constituent
of
natural
waters,
is
oxidized
by
hypochlorous
acid
(
HOCl
3)
to
form
hypobromous
acid
(
HOBr)
in
the
following
reaction:

Br­
+
HOCl
3

HOBr
+
Cl­

Hypobromous
acid
reacts
with
naturally
occurring
organic
substances
in
water
(
e.
g.,
humic
and
fulvic
acids)
to
yield
the
bromine­
containing
trihalomethanes
bromoform,
dibromochloromethane
and
bromodichloromethane
(
in
increasing
order
of
formation
rates)
(
Jolley
et
al.,
1978).

Trihalomethanes
may
also
be
produced
by
reaction
with
endogenous
organic
material
in
the
gut.
Mink
et
al.
(
1983)
treated
adult
male
Sprague­
Dawley
rats
with
a
single
oral
dose
of
48
mg
Cl
(
as
sodium
chloroacetate)
and
32
mg
Br­
(
as
potassium
bromide).
All
three
brominated
trihalomethanes
were
detected
in
the
stomach
contents
of
nonfasted
rats
following
treatment
(
Mink
et
al.,
1983).
Bromoform
and
dibromochloromethane
were
also
detected
in
the
plasma.

In
the
past,
bromoform,
bromodichloromethane
and
dibromochloromethane
have
been
used
in
pharmaceutical
manufacturing
and
chemical
synthesis,
as
ingredients
in
fire­
resistant
chemicals
and
gauge
fluids,
and
as
solvents
for
waxes,
greases,
resins,
and
oils
(
U.
S.
EPA,
1975).
However,
use
patterns
have
changed
over
time.
At
present,
the
primary
use
of
bromodichloromethane
is
as
a
chemical
intermediate
for
organic
synthesis
and
as
a
laboratory
reagent
(
ATSDR
1989).
Dibromochloromethane
is
reportedly
used
in
laboratory
quantities
only
(
ATSDR
1990).
Use
of
bromoform
is
limited
to
performance
of
geological
tests,
use
as
a
laboratory
reagent,
and
use
in
quality
assurance
programs
in
the
electronics
industry
(
ATSDR
1990).

B.
Summary
Brominated
trihalomethanes
are
volatile
organic
liquids
that
occur
in
drinking
water
as
byproducts
of
disinfection
with
chlorine.
The
brominated
trihalomethanes
occurring
in
water
are
bromoform,
dibromochloromethane
and
bromodichloromethane.
These
compounds
are
formed
in
water
when
hypochlorous
acid
oxidizes
bromide
ions
to
form
hypobromous
acid,
which
subsequently
reacts
with
organic
material.
In
the
past,
individual
brominated
trihalomethanes
have
been
used
for
a
variety
of
industrial
purposes.
Currently,
these
compounds
are
used
as
laboratory
reagents
and,
in
the
case
of
bromodichloromethane,
as
an
intermediate
in
chemical
synthesis.
Draft
­
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or
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February
20,
2003
III
­
1
III.
TOXICOKINETICS
This
section
summarizes
available
information
on
the
absorption,
distribution,
metabolism
and
excretion
of
brominated
trihalomethanes.
Because
the
toxicokinetic
properties
of
brominated
trihalomethanes
appear
to
be
generally
similar,
data
in
this
section
are
presented
for
this
class
of
compounds
as
a
group,
rather
than
by
individual
chemical.

A.
Absorption
Mink
et
al.
(
1986)
compared
the
absorption
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
male
Sprague­
Dawley
rats
and
male
B6C3F
1
mice.
The
study
animals
received
single
oral
doses
of
14C­
labeled
compound
in
corn
oil
by
gavage
at
dose
levels
of
100
mg/
kg
(
rats)
or
150
mg/
kg
(
mice).
Total
recovery
of
label
in
exhaled
air,
urine,
or
tissues
after
8
hours
ranged
from
62%
to
93%
(
Table
III­
1),
indicating
that
gastrointestinal
absorption
was
high
for
all
three
compounds.
The
level
of
radiolabeled
carbon
monoxide
in
exhaled
air
was
not
quantified
in
this
experiment.
Carbon
monoxide
has
since
been
recognized
as
a
product
of
brominated
trihalomethane
catabolism.

Mathews
et
al.
(
1990)
administered
14C­
bromodichloromethane
by
gavage
in
corn
oil
to
male
Fischer
344
rats
at
doses
of
1,
10,
32,
or
100
mg/
kg,
and
monitored
the
radiolabel
in
exhaled
air,
urine,
feces,
and
tissues.
Absorption
was
extensive,
with
at
least
86%
of
the
dose
recovered
as
expired
volatiles,
CO
2,
or
CO.
Only
small
amounts
were
recovered
in
urine
(<
5%)
or
in
feces
(<
3%)
within
24
hours
of
administration,
regardless
of
the
size
of
the
dose
(
Table
III­
2).

Table
III­
1
Recovery
of
Label
8
Hours
after
Oral
Administration
of
14C­
Labeled
Brominated
Trihalomethanes
to
Male
Sprague­
Dawley
Rats
or
Male
B6C3F1
Mice
Chemical
Percent
of
Label
Species
Expired
CO2
Expired
Parent
Urine
Organs
Total
Recovery
Bromodichloromethane
Rat
14.2
41.7
1.4
3.3
62.7
Mouse
81.2
7.2
2.2
3.2
92.7
Dibromochloromethane
Rat
18.2
48.1
1.1
1.4
70.3
Mouse
71.6
12.3
1.9
5.0
91.6
Bromoform
Rat
4.3
66.9
2.2
2.1
78.9
Mouse
39.7
5.7
4.6
12.2
62.2
Adapted
from
Mink
et
al.
(
1986)
and
U.
S.
EPA
(
1994b).
Draft
­
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or
Quote
February
20,
2003
III
­
2
Table
III­
2
Cumulative
Excretion
of
Label
after
Oral
Administration
of
14C­
Labeled
Bromodichloromethane
to
Male
F344
Rats
Dose
Time
(
hrs
posttreatment
Percent
of
Dose
Expired
CO2
Expired
CO
Expired
Volatiles
Urine
Feces
Total
Recovery
1
mg/
kg
1
9.5
±
1.1
NRa
2.1
±
1.5
NR
NR
11.6
±
1.3
4
37.0
±
3.2
1.5
±
0.7
2.7
±
1.8
NR
NR
41.1
±
2.8
8
62.9
±
2.2
2.7
±
1.1
NR
NR
NR
68.
±
1.7
16
76.4
±
3.2
NR
NR
NR
NR
81.8
±
2.9
24
77.5
±
3.3
3.3
±
1.5
3.0
±
1.6
4.1
±
0.2
2.7
±
1.5
90.7
±
1.8
10
mg/
kg
1
8.0
±
2.0
NR
2.0
±
0.8
NR
NR
10.0
±
1.85
4
39.9
±
3.2
1.9
±
0.4
2.7
±
1.1
NR
NR
44.5
±
3.0
8
66.0
±
4.0
3.4
±
0.9
NR
NR
NR
72.1
±
3.9
16
81.3
±
1.7
NR
NR
NR
NR
87.4
±
1.5
24
82.1
±
1.8
4.3
±
1.0
2.8
±
1.1
4.3
±
0.2
0.7
±
0.2
94.2
±
1.6
100
mg/
kg
1
1.9
±
0.9
0.1
±
0
1.5
±
1.2
NR
NR
4.6
±
1.8
4
5.5
±
1.8
0.3
±
0.1
4.2
±
1.9
NR
NR
10.0
±
2.9
8
NR
NR
NR
0.6
±
0.4
NR
10.6
±
3.0
16
33.4
±
7.4
2.3
±
0.7
5.7
±
2.1
NR
NR
42.0
±
8.3
24
71.0
±
1.7
5.2
±
0.3
6.3
±
2.1
4.1
±
0.2
0.7
±
0.3
87.3
±
1.6
Adapted
from
Mathews
et
al.
(
1990)
and
U.
S.
EPA
(
1994b).
aNot
reported
Draft
­
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or
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February
20,
2003
III
­
3
Lilly
et
al.
(
1998)
examined
the
effects
of
vehicle
on
the
absorption
of
orally
administered
bromodichloromethane
in
an
experiment
designed
to
develop
and
validate
a
physiologically­
based
pharmacokinetic
model.
Male
F344
rats
(
3
animals/
dose/
vehicle/
assay)
were
gavaged
with
0,
50,
or
100
mg
bromodichloromethane/
kg
in
either
corn
oil
or
10%
Emulphor
®
,
and
bromodichloromethane
levels
were
monitored
in
blood
and
exhaled
air.
The
dose
levels
approximated
doses
previously
utilized
in
two­
year
cancer
bioassays
of
bromodichloromethane
(
NTP,
1987).
Concentrations
of
bromodichloromethane
in
blood
and
exhaled
air
peaked
rapidly,
reaching
maximal
concentrations
less
than
one
hour
after
administration.
The
vehicle
of
administration
had
significant
effects
on
the
blood
and
exhaled
air
concentrations.
Delivery
of
bromodichloromethane
in
10%
Emulphor
®
resulted
in
faster
initial
uptake,
as
inferred
from
higher
blood,
tissue
and
breath
chamber
concentrations,
when
compared
to
corn
oil
(
data
presented
graphically).
At
6
hours
after
administration,
more
than
90%
and
100%
of
the
administered
dose
had
been
absorbed
from
the
corn
oil
and
Emulphor
®
vehicles,
respectively.

B.
Distribution
Data
on
the
distribution
of
brominated
trihalomethanes
in
exposed
humans
are
limited.
Roth
(
1904)
measured
the
bromoform
content
of
tissues
of
a
man
who
died
from
an
accidental
oral
overdose
of
bromoform
and
found
levels
in
stomach
and
lung
of
130
and
90
mg/
kg
wet
weight,
respectively.
Lower
levels
were
reported
in
the
intestine,
liver,
kidney,
and
brain.
Pellizzari
et
al.
(
1982)
measured
trihalomethanes
in
42
samples
of
human
milk
taken
from
women
in
urban
areas.
Dibromochloromethane
was
detected
in
one
sample.
Neither
the
level
measured
nor
the
detection
limit
were
reported
for
this
study.

Data
on
the
distribution
of
brominated
trihalomethanes
in
animals
are
available
from
studies
in
rats
and
mice.
Mink
et
al.
(
1986)
compared
the
distribution
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
male
Sprague­
Dawley
rats
and
male
B6C3F
1
mice.
Single
oral
doses
of
14C­
labeled
compound
in
corn
oil
were
administered
by
gavage
at
dose
levels
of
100
mg/
kg
(
rats)
or
150
mg/
kg
(
mice).
Tissue
levels
of
radioactivity
were
measured
8
hours
after
dose
administration.
The
chemical
form
of
the
label
measured
in
the
tissues
(
e.
g.
parent
or
metabolite,
bound
or
free)
was
not
determined.
In
the
rat,
the
total
organ
content
of
label
ranged
from
1.4%
to
3.6%
for
the
various
compounds.
The
stomach,
liver,
and
kidneys
contained
higher
levels
than
bladder,
brain,
lung,
muscle,
pancreas,
and
thymus.
In
mice,
4%
to
5%
of
the
administered
compound
was
recovered
in
the
organs.
However,
an
additional
10%
of
the
label
associated
with
bromoform
was
recovered
in
the
blood
of
mice,
yielding
total
organ
levels
of
12%
to
14%.
The
authors
attributed
this
elevated
recovery
of
label
to
carboxyhemoglobin
formation.
The
levels
of
carboxyhemoglobin
were
not
measured
in
this
experiment.

Mathews
et
al.
(
1990)
investigated
the
distribution
of
bromodichloromethane
following
oral
exposure
in
male
Fischer
344
rats.
Animals
were
given
a
single
oral
gavage
dose
of
1,
10,
32,
or
100
mg/
kg
of
14C­
bromodichloromethane
dissolved
in
corn
oil.
Approximately
3%
to
4%
of
the
administered
dose
was
detected
in
tissues
after
24
hours.
The
highest
levels
(
1%
to
3%)
were
measured
in
liver.
Repeated
doses
of
10
or
100
mg/
kg­
day
for
10
days
resulted
in
total
retention
of
only
0.9%
to
1.1%
of
the
administered
label,
and
had
no
effect
on
the
tissue
Draft
­
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or
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February
20,
2003
III
­
4
distribution
of
bromodichloro­
methane.

The
Chlorine
Chemistry
Council
(
CCC,
2000c)
sponsored
a
study
which
analyzed
the
levels
of
bromodichloromethane
in
parental
tissues
and
fluids
and
F
1
generation
tissues
as
part
of
a
reproductive
and
developmental
study
in
Sprague­
Dawley
rats
(
see
Section
V.
E.
1
for
a
full
study
description).
Data
from
this
study
were
summarized
in
Christian
et
al.
(
2001b).
Bromodichloromethane
was
administered
in
the
drinking
water
at
concentrations
of
0,
50,
150,
450,
or
1350
ppm.
The
estimated
dosage
on
a
mg/
kg­
day
basis
varied
with
the
stage
of
the
study
(
see
Table
V­
7).
Plasma
and
other
tissue
samples
were
collected
for
analysis
as
described
in
Table
III­
3.
All
samples
were
maintained
frozen
and
shipped
to
the
analytical
lab
(
Lancaster
Laboratories,
Lancaster,
PA).
Analysis
of
plasma
collected
from
male
and
female
rats
prior
to
mating
and
from
female
rats
during
gestation
and
lactation
did
not
detect
quantifiable
levels
of
bromodichloromethane
(
limit
of
detection
0.11
µ
g).
Bromodichloromethane
was
detected
at
a
concentration
of
0.38
µ
g/
g
in
the
milk
from
one
female
in
the
1350
ppm
group.
Bromodichloromethane
was
not
detected
in
placentas,
amniotic
fluid,
or
fetal
tissue
collected
on
GD
21
or
in
plasma
collected
from
postpartum
day
29
weanling
pups.

Table
III­
3
Over
view
of
Tissue
Collection
for
Analysis
of
Bromodichloromethane
in
Sprague­
Dawley
Rat
Tissues
and
Fluids
(
CCC,
2000c).

Generation
Sex
Physiological
state
Tissue
Day
of
collection
No.
of
Samples;
Collection
freq.
Comments
P
M,
F
Pre­
mating
Plasma
Day
1
of
exposure
3
rats/
sex/
group;
3
times/
day
­

P
M,
F
Pre­
mating
Plasma
Day
14
of
exposure
3
rats/
sex/
group;
3
times/
day
­

P
F
Pregnant
Plasma
GD
20
3
rats/
group;
3
times/
day
Rats
continuously
exposed
since
study
day
1
P
F
Pregnant
Placenta
amniotic
fluid,
and
fetuses
GD
21
3
litters/
day
Tissues
pooled
by
litter
P
F
Lactating
Plasma
LD
15
3
rats/
group
3
times/
day
P
F
Lactating
Milk
LD
15
3
rats/
group
1
IU
oxytocin
admin.
by
IV
approx.
5
min.
before
milking
F1
M,
F
Weaning
Plasma
LD
29
3
pups/
sex;
3
litters;
3
times/
day
­

Modified
from
CCC
(
2000c)
Draft
­
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Not
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or
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February
20,
2003
III
­
5
Abbreviations:
P,
parental;
M,
male;
F,
female;
GD,
gestation
day;
LD,
lactation
day
The
Chlorine
Chemistry
Council
(
CCC,
2000a)
sponsored
a
study
which
analyzed
the
levels
of
bromodichloromethane
in
parental
tissues
and
fluids
and
F
1
generation
tissues
as
part
of
a
reproductive
and
developmental
study
in
New
Zealand
White
rabbits
(
see
Section
V.
E.
1
for
a
full
study
description).
Data
from
this
study
were
summarized
in
Christian
et
al.
(
2001b).
Bromodichloromethane
was
administered
to
groups
of
rabbits
(
4/
concentration)
in
the
drinking
water
at
concentrations
of
0,
50,
150,
450,
or
1350
ppm.
The
estimated
doses
at
these
concentrations
were
0,
4.9,
13.9,
32.3,
or
76.3
mg/
kg­
day,
respectively.
Blood
samples
were
collected
on
GD
7
and
28.
Amniotic
fluid,
and
placenta
samples
were
collected
on
GD
29
after
collection
of
a
third
blood
sample,
and
amniotic
fluid
and
placental
tissue
were
pooled
by
litter.
Blood
samples
were
collected
from
three
randomly
selected
fetuses
per
litter.
Bromodichloromethane
was
detected
at
concentrations
of
0.15
and
0.17
µ
g/
g
(
limit
of
detection
0.11
µ
g/
g)
in
placentas
from
two
litters
in
the
1350
ppm
exposure
group.
Bromodichloromethane
was
detected
in
one
fetus
from
the
1350
ppm
group
"...
at
a
level
below
the
limit
of
detection".
Bromodichloromethane
was
not
detected
in
placentas
from
does
exposed
to
concentrations
up
to
450
ppm,
in
amniotic
fluid
from
does
exposed
to
concentrations
up
to
1350
ppm,
or
in
the
remaining
fetuses
of
does
exposed
to
concentrations
as
high
as
1350
ppm.

C.
Metabolism
1.
Overview
The
toxicity
of
the
brominated
trihalomethanes
is
mediated
by
cytochrome
P450­
mediated
bioactivation
to
reactive
metabolites.
The
pathways
for
brominated
trihalomethane
metabolism
were
initially
inferred
from
studies
of
the
structurally­
related
trihalomethane
chloroform
(
U.
S.
EPA,
1994b).
Additional
details
of
brominated
trihalomethane
metabolism
have
subsequently
been
elucidated
in
a
number
of
laboratories
using
both
in
vitro
and
in
vivo
approaches
and
are
described
below.
Figure
III­
1
presents
a
general
metabolic
scheme
for
chloroform
and
the
brominated
trihalomethanes.

The
metabolism
of
brominated
trihalomethanes
occurs
via
at
least
two
pathways
(
U.
S.
EPA
1994b).
The
oxidative
pathway
requires
NADPH
and
oxygen,
whereas
the
reductive
pathway
can
utilize
NADPH
or
NADH
and
is
inhibited
by
oxygen.
Both
reactions
are
believed
to
be
mediated
by
cytochrome
P450
isoforms.
In
the
presence
of
oxygen
(
oxidative
metabolism),
the
reaction
product
is
trihalomethanol
(
CX
3
OH),
which
spontaneously
decomposes
to
yield
a
reactive
dihalocarbonyl
(
CX
2
O)
such
as
phosgene
(
CCl
2
O).
Dihalocarbonyls
may
undergo
a
variety
of
reactions,
such
as
adduct
formation
with
various
cellular
nucleophiles,
hydrolysis
to
yield
carbon
dioxide,
or
glutathione­
dependent
reduction
to
yield
carbon
monoxide.
When
oxygen
levels
are
low
(
reductive
metabolism),
the
metabolic
reaction
products
appear
to
be
free
radical
species
such
as
the
dihalomethyl
radical
(°
CHX
2).
These
radicals
are
highly
reactive
and
may
also
form
covalent
adducts
with
a
variety
of
cellular
molecules.
Evidence
supporting
this
metabolic
scheme
and
information
on
species
differences
in
the
rate
and
extent
of
trihalomethane
metabolism
are
presented
below.
Additional
data
derived
from
studies
of
chloroform
Draft
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III
­
6
CHX3
CX3OH
CX2O
CHX2
P­
450
P­
450
X­
NADPH
or
NADH
2NADPH
O2
H2O
HX
Oxidative
Pathway
Reductive
Pathway
CO+
GSSG
CO2
OTZ
RCX2OH
RH
2GSH
H2O
Cysteine
2HX
2HX
2HX
Figure
III­
1
Proposed
Metabolic
Pathways
for
Brominated
Trihalomethanes
X
=
halogen
atom
(
chlorine
or
bromine);
R
=
cellular
nucleophile
(
protein,
nucleic
acid);

GSH
=
reduced
glutathione;
GSSG
=
oxidized
glutathione;

OTZ
=
oxothiazolidine
carboxylic
acid;
P­
450
=
cytochrome
P­
450
Adapted
from
Stevens
and
Anders
(
1981);
Tomasi
et
al.
(
1985)
Draft
­
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or
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20,
2003
III
­
7
are
described
in
U.
S.
EPA
(
1994b).

The
metabolism
of
trihalomethanes
(
including
chloroform
and
the
brominated
trihalomethanes)
has
been
most
intensively
studied
using
chloroform
as
a
substrate.
These
studies
indicate
that
many
factors
influence
metabolism,
including
strain,
species,
chloroform
concentration,
and
possibly
gender.
A
comprehensive
review
of
chloroform
studies
is
beyond
the
scope
of
this
document.
However,
because
the
P450­
mediated
metabolism
of
the
brominated
trihalomethanes
is
expected
to
be
similar
to
that
of
chloroform,
descriptions
of
a
few
representative
studies
of
chloroform
metabolism
are
provided
to
provide
additional
background
information
on
the
metabolism
of
trihalomethanes.

A
key
question
in
hazard
characterization
is
the
identity
of
the
P450
isoforms
responsible
for
bioactivation
of
brominated
trihalomethanes
to
reactive
metabolites.
This
is
because
individuals
or
subpopulations
with
elevated
levels
of
these
enzymes
may
be
at
greater
risk
for
adverse
effects.
The
identities
of
the
cytochrome
P450
isoforms
responsible
for
trihalomethane
metabolism
have
been
investigated
most
intensively
in
studies
of
chloroform
(
studies
of
brominated
trihalomethanes
are
described
in
sections
2
and
3
below).

Studies
by
Nakajima
et
al.
(
1995)
and
Testai
et
al.
(
1996)
indicate
that
chloroform
concentration
plays
a
critical
role
in
determining
the
role
of
different
isoforms
and
the
associated
effects
of
metabolic
inducers.
Nakajima
et
al.
(
1995)
pretreated
male
Wistar
rats
with
three
inducers
of
specific
P450
isoforms
and
subsequently
administered
a
single
dose
of
chloroform
by
gavage
in
corn
oil.
The
inducers
used
were
phenobarbital
(
CYP2B1/
2),
n­
hexane
(
CYP2E1),
and
2­
hexanone
(
CYP2B1/
2
and
CYP2E1).
Liver
damage
(
as
determined
by
serum
enzyme
activity
and
histopathology)
was
greatest
at
the
mid­
dose
in
the
hexane­
treated
animals.
In
contrast,
rats
pretreated
with
phenobarbital
or
2­
hexanone
showed
a
dose­
related
increase
of
liver
damage
at
all
dose
levels.
The
pattern
of
damage
was
consistent
in
each
case
with
the
tissue
distribution
patterns
of
the
induced
cytochrome
P450
isoform(
s).
The
study
authors
concluded
on
the
basis
of
these
results
that
CYP2E1
catalyzes
chloroform
metabolism
at
low
doses
and
that
CYP2B1/
2
catalyzes
chloroform
metabolism
at
higher
doses.

While
experimental
evidence
indicates
that
CYP2E1
and
CYPB1/
2
catalyze
the
oxidative
pathway,
the
identities
of
the
cytochrome
P450
isoforms
which
catalyze
the
reductive
pathway
have
not
been
established.
In
general,
CYP2E1
protein
can
catalyze
reductive
as
well
as
oxidative
reactions
(
Lieber,
1997)
and
this
isoform
has
been
implicated
in
the
production
of
trichloromethyl
radicals
from
carbon
tetrachloride
(
see
Lieber
et
al.
1997).
However,
evidence
for
a
dual
role
of
either
CYP2E1
or
CYP2B1/
2
in
catalyzing
the
oxidative
and
reductive
pathways
for
trihalomethane
metabolism
has
been
contradictory,
perhaps
as
a
result
of
the
different
concentrations
of
chloroform
used
in
different
experiments
(
summarized
in
Testai
et
al.
1996).
To
address
the
issue
of
concentration,
Testai
et
al.
(
1996)
studied
the
role
of
different
isoforms
in
chloroform
using
microsomes
prepared
from
Sprague­
Dawley
rats
pretreated
with
a
variety
of
cytochrome
P450
inducers.
The
microsomes
were
incubated
under
varying
conditions
of
chloroform
concentration,
oxygenation,
and
presence
of
isoform­
specific
inhibitors
or
antibodies.
Under
the
conditions
utilized
in
this
series
of
experiments,
the
authors
concluded
that
the
Draft
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February
20,
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III
­
8
cytochrome
P450
isoforms
involved
in
oxidative
metabolism
of
brominated
trihalomethanes
do
not
participate
in
the
reductive
pathway.

Recent
evidence
from
studies
in
Salmonella
typhimurium
strains
engineered
to
express
the
rat
glutathione
S­
transferase
theta
1­
1
(
GSTT1­
1)
gene
suggests
that
bioactivation
of
brominated
trihalomethanes
to
mutagenic
species
is
also
mediated
by
one
or
more
glutathione
S­
transferasemediated
conjugation
pathways
(
Pegram
et
al.,
1997;
DeMarini
et
al.,
1997).
Details
of
these
studies
are
presented
in
Section
V.
E.
Proposed
routes
for
glutathione
conjugation
and
bioactivation
are
illustrated
in
Figure
V­
2,
located
in
Section
V.
E.

2.
In
Vitro
Studies
Ahmed
et
al.
(
1977)
investigated
the
in
vitro
oxidative
(
aerobic)
metabolism
of
brominated
trihalomethanes
to
carbon
monoxide
by
the
rat
liver
microsomal
fraction.
Metabolism
of
bromoform
resulted
in
the
highest
level
of
carbon
monoxide
formation,
followed
by
dibromochloromethane
and
bromodichloromethane
in
decreasing
order.
Glutathione,
NADPH
and
oxygen
were
required
for
maximal
carbon
monoxide
production.
This
activity
was
inducible
by
phenobarbital
or
3­
methylcholanthrene
pretreatment
(
agents
which
are
known
to
increase
cytochrome
P­
450
activity)
and
was
inhibited
by
the
cytochrome
P­
450
inhibitor
SKF
525­
A.
Similar
results
were
reported
by
Stevens
and
Anders
(
1979).
In
addition,
Stevens
and
Anders
(
1979)
reported
the
formation
of
2­
oxothiazolidine­
4­
carboxylic
acid
(
OTZ)
when
bromoform
was
incubated
in
the
presence
of
cysteine.
Dihalocarbonyls
react
with
cysteine
to
form
OTZ.
Thus,
detection
of
OTZ
provides
evidence
that
a
dihalocarbonyl
intermediate
was
formed
during
bromoform
metabolism.

Wolf
et
al.
(
1977)
studied
the
in
vitro
metabolism
of
bromoform
and
chloroform
to
carbon
monoxide
under
anaerobic
conditions
using
liver
preparations
from
phenobarbital­
induced
rats.
Bromoform
metabolism
resulted
in
much
greater
levels
of
carbon
monoxide
production
than
did
the
metabolism
of
chloroform.
Gao
and
Pegram
(
1992)
reported
that
binding
of
reactive
intermediates
to
rat
hepatic
microsomal
lipid
and
protein
under
reductive
(
anaerobic)
conditions
was
more
than
twice
as
high
for
bromodichloromethane
as
for
chloroform.
These
data
suggest
that
reductive
metabolism
may
be
a
more
important
pathway
for
metabolism
of
brominated
trihalomethanes
than
for
chloroform.

Tomasi
et
al.
(
1985)
studied
the
anaerobic
activation
of
bromoform,
bromodichloromethane,
and
chloroform
to
free
radical
intermediates
in
vitro
using
rat
hepatocytes
isolated
from
phenobarbital­
induced
male
Wistar
rats.
The
production
of
a
free
radical
intermediate
was
measured
by
electron
spin
resonance
(
ESR)
spectroscopy
using
the
spin
trap
compound
phenyl­
t­
butylnitrone.
The
intensity
of
the
ESR
signal
was
greatest
for
bromoform,
followed
by
bromodichloromethane
and
then
chloroform.
The
largest
ESR
signal
was
detected
when
hepatocytes
were
incubated
under
anaerobic
conditions.
Incubation
in
the
presence
of
air
resulted
in
a
reduction
of
the
signal,
as
did
addition
of
cytochrome
P­
450
inhibitors
such
as
SKF­
525A,
metyrapone,
and
carbon
monoxide.
These
data
were
interpreted
to
indicate
that
freeradical
formation
depended
on
reductive
metabolism
of
the
trihalomethanes
mediated
by
the
Draft
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February
20,
2003
III
­
9
cytochrome
P450
system.
Comparison
of
the
ESR
spectra
for
chloroform,
deuterated
chloroform,
and
bromodichloromethane
indicated
that
the
free
radical
intermediate
produced
by
chloroform
metabolism
was
°
CHCl
2.
The
authors
speculated
that
the
brominated
trihalomethanes
are
also
metabolized
by
transfer
of
an
electron
directly
from
the
cytochrome
to
the
halocompound
with
the
successive
formation
of
the
dihalomethyl
radical
(°
CHX
2)
and
a
halide
ion
(
X­).

3.
In
Vivo
Studies
Mink
et
al.
(
1986)
compared
the
metabolic
products
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
male
Sprague­
Dawley
rats
and
male
B6C3F
1
mice
(
strain
not
reported).
Animals
were
given
a
single
oral
dose
of
14C­
labeled
compound
by
gavage
in
corn
oil
at
dose
levels
of
100
mg/
kg
for
rats
and
150
mg/
kg
for
mice.
Levels
of
14C
were
measured
in
exhaled
carbon
dioxide
recovered
within
8
hours
after
dose
administration.
Expired
carbon
dioxide
accounted
for
4.3%
to
18.2%
of
the
administered
label
in
rats
(
Table
III­
1),
suggesting
that
the
parent
compound
had
undergone
limited
metabolism
and
oxidation.
In
mice,
the
fraction
of
label
excreted
as
carbon
dioxide
was
higher,
ranging
from
40%
to
81%.
These
data
indicate
that
oxidative
metabolism
of
brominated
trihalomethanes
to
carbon
dioxide
was
more
rapid
and
extensive
(
by
a
factor
of
four­
to
ninefold)
in
mice
than
in
rats.
As
previously
noted
in
Section
III.
A,
production
of
carbon
monoxide,
a
known
metabolite
of
brominated
trihalomethanes,
was
not
measured
in
this
study.

Anders
et
al.
(
1978)
investigated
the
formation
of
carbon
monoxide
from
brominated
trihalomethanes
in
corn
oil
administered
to
Sprague­
Dawley
rats
at
doses
of
1
mmol/
kg
(
119
to
252
mg/
kg)
by
intraperitoneal
injection.
Bromoform
produced
the
highest
levels
of
blood
carbon
monoxide,
followed
by
dibromochloromethane
and
bromodichloromethane
in
decreasing
order.
A
dose­
response
relationship
was
noted
for
bromoform
following
administration
of
252,
506,
or
1,012
mg/
kg.
Carbon
monoxide
production
was
inducible
by
pretreatment
with
phenobarbital,
but
pretreatment
with
3­
methylcholanthrene
had
no
effect.
Carbon
monoxide
production
was
significantly
inhibited
by
SKF­
525­
A.
Administration
of
3H­
bromoform
resulted
in
decreased
carbon
monoxide
formation
when
compared
to
bromodichloromethane
and
dibromochloromethane,
indicating
that
the
carbon­
hydrogen
bond
breakage
may
be
the
ratelimiting
step
under
aerobic
conditions.
Similar
results
were
later
reported
by
Stevens
and
Anders
(
1981).

Tomasi
et
al.
(
1985)
studied
the
in
vivo
metabolism
of
chloroform,
bromodichloromethane
and
bromoform
to
free
radical
intermediates
in
rats.
Starved,
phenobarbital­
induced
male
Wistar
rats
(
number
not
stated)
were
given
intraperitoneal
injections
of
1,100
mg/
kg
chloroform,
820
mg/
kg
bromodichloromethane,
or
1,260
mg/
kg
bromoform
dissolved
in
olive
oil.
The
animals
were
sacrificed
and
the
livers
were
homogenized.
The
production
of
a
free
radical
intermediate
by
the
livers
was
determined
by
ESR
spectroscopy.
The
authors
reported
detection
of
free
radicals
in
the
livers
of
all
treated
rats.
The
intensity
of
the
ESR
signal
followed
a
ranking
similar
to
that
observed
in
in
vitro
experiments
(
bromoform
>
bromodichloromethane
>
chloroform),
confirming
that
the
reductive
formation
of
free
radicals
is
greater
for
brominated
trihalomethanes
than
for
chloroform.
Draft
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20,
2003
III
­
10
Mathews
et
al.
(
1990)
studied
the
metabolism
of
14C­
bromodichloromethane
in
male
Fischer
344
rats.
Animals
(
n
=
4)
were
given
a
single
oral
dose
of
1,
10,
32,
or
100
mg/
kg
of
bromodichloromethane
dissolved
in
corn
oil.
Levels
of
labeled
carbon
dioxide
and
carbon
monoxide
in
exhaled
air
were
measured
for
24
hours.
Approximately
70%
to
80%
of
the
dose
was
metabolized
and
exhaled
as
14CO
2
and
3%
to
5%
of
the
dose
as
14CO.
However,
14CO
2
production
was
slower
following
a
single
dose
of
100
mg/
kg
than
after
the
administration
of
a
single
dose
of
32
mg/
kg
or
lower,
suggesting
saturation
of
metabolism.
Repeated
doses
of
100
mg/
kg­
day
for
10
days
resulted
in
an
increased
rate
of
14CO
2
production,
compared
with
the
initial
rate.
The
authors
concluded
on
the
basis
of
these
data
that
bromodichloromethane
may
induce
its
own
metabolism.

Thornton­
Manning
et
al.
(
1994)
evaluated
the
effect
of
bromodichloromethane
exposure
on
cytochrome
P450
isozyme
activity
in
female
F344
rats
(
6/
dose).
Gavage
doses
of
75
to
300
mg/
kg­
day
were
administered
in
a
solution
of
10%
Emulphor
®
(
an
emulsifier)
for
five
consecutive
days.
Treatment
resulted
in
decreased
activity
of
the
CYP1A
and
CYP2B
isozymes.
In
contrast,
there
was
no
effect
on
CYP2E1
activity.

Pankow
et
al.
(
1997)
investigated
the
metabolism
of
dibromochloromethane
in
male
Wistar
rats
following
single
and
repeated
gavage
doses.
Rats
receiving
a
single­
dose
(
6
animals
per
group)
were
treated
with
0
(
vehicle
only),
0.4,
0.8,
1.6
or
3.1
mmol
dibromochloromethane
kg
dissolved
in
olive
oil.
Rats
receiving
multiple
doses
were
gavaged
with
0
(
vehicle
only)
or
0.8
mmol
dibromochloromethane/
kg
once
a
day
for
7
days.
The
blood
or
plasma
concentrations
of
parent
compound,
bromide,
and
carbon
monoxide
(
as
carboxyhemoglobin,
COHb)
were
measured
following
dibromochloromethane
administration.
The
level
of
oxidized
glutathione
(
GSSG)
in
the
liver
of
treated
animals
was
also
assayed.
Oral
administration
of
dibromochloromethane
resulted
in
a
significant
elevation
of
plasma
bromide
levels
at
all
doses
tested.
Bromide
did
not
return
to
baseline
levels
even
after
10
days.
Repeated
administration
of
0.8
mmol
dibromochloromethane/
kg
resulted
in
significantly
higher
plasma
levels
of
bromide
than
were
measured
following
a
single
dose
of
0.8
mmol/
kg.
COHb
was
also
elevated
in
a
dosedependent
manner
following
either
single
or
repeated
administration
of
dibromochloromethane,
but
returned
to
baseline
levels
within
24
hours
after
treatment.
GSSG
levels
were
significantly
increased
at
12­
and
24­
hour
time
points
following
a
single
0.8
mmol/
kg
dose
(
no
other
doses
were
examined).
Levels
of
GSSG
returned
to
baseline
levels
by
48
hours
after
treatment.

Pankow
et
al.
(
1997)
conducted
additional
experiments
to
determine
whether
reduced
glutathione
(
GSH)
is
a
requirement
for
in
vivo
dibromochloromethane
metabolism
and
to
identify
P450
isozymes
involved
in
the
metabolism
of
dibromochloromethane.
Pretreatment
of
rats
with
buthionine
sulfoximine
(
an
agent
which
depletes
GSH)
reduced
GSH
concentrations
as
anticipated
and
decreased
the
rate
of
bromide
and
COHb
production.
In
contrast,
pretreatment
with
butylated
hydroxyanisole
(
which
increases
GSH
levels)
increased
the
rate
of
bromide
and
COHb
production.
These
results
suggest
that
GSH
plays
a
role
in
dibromochloromethane
metabolism.
Further
studies
were
conducted
to
determine
which
cytochrome
P450
isoform(
s)
participate
in
the
in
vivo
metabolism
of
dibromochloromethane.
Simultaneous
exposure
to
0.8
Draft
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February
20,
2003
III
­
11
mmol
dibromochloromethane/
kg
and
diethyldithiocarbamate
(
a
potent
inhibitor
of
P450
isoform
CYP2E1)
partially
inhibited
the
production
of
bromide
and
COHb.
In
contrast,
pretreatment
with
isoniazid
(
an
potent
inducer
of
CYP2E1)
increased
formation
of
bromide
and
COHb.
These
experiments
indicate
that
CYP2E1
is
at
least
partially
responsible
for
dibromochloromethane
metabolism.
Pretreatment
with
phenobarbital,
an
inducer
of
cytochrome
P450
isoforms
CYP2B1
and
2B2,
increased
the
concentration
of
bromide
in
plasma,
suggesting
that
CYP2B1
and
2B2
may
also
participate
in
the
catabolism
of
dibromochloromethane.
Pretreatment
with
m­
xylene,
which
induces
both
CYP2E1
and
CYP2B1/
2,
resulted
in
higher
bromide
levels
than
inducers
of
CYP2E1
(
isoniazid)
or
CYP2B1/
2
(
phenobarbital)
administered
individually.
Pankow
et
al.
(
1997)
concluded
on
the
basis
of
these
multiple
experiments
that
1)
bromide
and
carbon
monoxide
are
metabolites
of
dibromochloromethane;
2)
dibromochloromethane
is
metabolized
via
the
oxidative
pathway;
3)
the
oxidative
metabolism
of
dibromochloromethane
is
catalyzed
by
CYP2E1
and
CYP2B1/
2;
and
4)
dibromochloromethane
plays
a
role
in
the
induction
of
CYP2E1.

Allis
et
al.
(
2001)
investigated
the
effect
of
inhalation
exposure
to
bromodichloromethane
on
the
activity
and
protein
levels
of
CYP1A2,
CYP2B1,
and
CYP2E1
in
female
F344
rats
(
6/
dose).
In
addition,
the
effect
of
inhalation
exposure
on
the
activity
level
of
CYP1A1
was
assayed.
Serum
bromide
ion
concentration,
an
indicator
of
the
total
metabolism
of
bromodichloromethane,
was
measured
in
samples
drawn
from
a
separate
set
of
animals
(
4/
concentration)
exposed
to
the
same
concentrations.
The
test
animals
were
exposed
for
4
hours
to
measured
bromodichloromethane
concentrations
of
0,
106,
217,
419,
812,
1620,
and
3240
ppm.
The
microsomal
isozyme
activities
assayed
were:
p­
nitrophenol
hydrolase
(
PNP),
an
indicator
of
CYP2E1
activity;
pentoxyresorufin­
O­
dealkylase
(
PROD),
an
indicator
of
CYP2B1/
2
activity;
ethoxsyresorufin­
O­
dealkylase
(
EROD),
an
indicator
of
CYP1A1
activity;
and
methoxyresorufin­
O­
dealkylase
(
MROD),
an
indicator
of
CYP1A2
activity.
The
pattern
of
results
for
isozyme
activity
obtained
in
this
inhalation
study
was
similar
to
that
reported
for
male
F344
rats
treated
with
bromodichloromethane
by
gavage.
CYP2E1
activity
as
measured
by
PNP
activity
was
not
significantly
affected
by
treatment.
MROD,
EROD,
and
PROD
activities
showed
modest
increases
at
low
exposure
concentrations.
The
increases
were
statistically
significant
for
EROD
and
MROD
at
the
106
ppm
exposure
concentration.
Decreases
were
observed
at
higher
exposure
concentrations
relative
to
controls.
These
decreases
were
statistically
significant
for
PROD
at
3240
ppm
and
for
EROD
and
MROD
at
concentrations
of
800
ppm
and
greater.
The
results
for
isozyme
protein
levels,
as
measured
by
Western
blots,
were
generally
consistent
with
the
results
for
isozyme
activity.
The
study
authors
speculated
that
the
most
dramatic
reductions
in
isozyme
activity
(
PROD
and
MROD)
were
a
result
of
suicide
inhibition.
In
addition,
they
concluded
that
analyses
of
the
EROD
and
MROD
activity
and
protein
level
patterns
indicates
that
CYP1A2
is
involved
in
the
metabolism
of
bromodichloromethane.
The
study
authors
noted
that
it
is
not
typical
for
this
isozyme
to
metabolize
the
small
molecules
(
such
as
chloroform)
that
are
the
usual
substrates
for
CYP2E1,
but
observed
that
the
presence
of
the
large
bromide
ion
may
make
bromodichloromethane
a
suitable
substrate
for
CYP1A2.
Blood
bromide
concentration
reached
a
maximum
at
200
ppm,
indicating
that
metabolism
was
saturated
a
concentrations
equal
to
or
greater
than
200
ppm.
Draft
­
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February
20,
2003
III
­
12
Allis
et
al.
(
2002)
reported
additional
evidence
for
metabolism
of
bromodichloromethane
by
CYP1A2.
Induction
of
CYP1A2
without
parallel
induction
of
CYP2E1
or
CYP2B1/
2
(
accomplished
by
pretreatment
with
2,3,7,8­
tetrachlorodibenzodioxin,
TCDD),
increased
hepatotoxicity
in
male
F344
rats
administered
a
gavage
dose
of
400
mg/
kg
bromodichloromethane.
Hepatotoxicity
was
assessed
by
measurement
of
alanine
aminotransferase
(
ALT)
and
sorbitol
dehydrogenase
(
SDH)
activity.
Pretreatment
with
TCDD
increased
serum
bromide
levels
(
a
measure
of
total
bromodichloromethane
metabolism)
in
rats
treated
with
200
or
400
mg/
kg
when
compared
to
uninduced
controls.
The
apparent
inconsistency
between
lack
of
hepatotoxicity
and
increased
total
metabolism
at
200
mg/
kg
was
explained
by
effective
detoxification
at
this
dose,
presumably
by
glutathione.
Selective
inhibition
of
CYP1A2
activity,
by
administration
of
isosafrole
to
TCDD­
induced
animals
prior
to
treatment
with
400
mg/
kg
bromodichloromethane
significantly
reduced
the
hepatotoxic
response
and
serum
bromide
concentrations.

Allis
and
colleagues
(
Allis
et
al.,
2002;
Allis
and
Zhao,
2002;
Zhao
and
Allis,
2002)
assessed
the
ability
of
various
rat
and
human
CYP
isoenzymes
to
metabolize
bromodichloromethane
and
determined
kinetic
parameters
for
those
showing
measurable
metabolic
activity.
Allis
and
Zhao
(
2002)
tested
five
rat
and
six
human
CYP
isoenzymes
in
vitro
for
metabolism
of
bromodichloromethane
using
recombinant
systems
expressing
single
isozyme
activities.
The
tested
recombinant
isoenzymes
were
rat
CYP2E1,
CYP2B1/
2,
CYP1A2,
CYP2C11,
AND
CYP3A1
and
human
CYP2E1,
CYP1A2,
CYP2A6,
CYP2B6,
CYP2D6
and
CYP3A4.
The
results
of
this
study
indicate
that
the
principal
metabolizing
enzymes
in
rat
are
those
identified
previously,
namely
CYP2E1,
CYP2B1/
2,
CYP1A2.
Results
for
CYP3A1
suggest
that
it
may
have
weak
metabolic
activity,
but
the
level
of
activity
was
not
sufficient
for
a
quantitative
assessment.
CYP2C11
was
not
active.
Human
CYP2E1,
CYP1A2,
and
CYP3A4
showed
substantial
metabolic
activity
toward
bromodichloromethane.
Human
CYP2A6
showed
lower,
but
measurable,
levels
of
activity.
CYP2B6
and
CYP2D6
were
not
active.
Based
on
these
assays,
only
CYP2E1
and
CYP1A2
metabolize
bromodichloromethane
in
both
species.
CYP2E1
is
the
high
affinity
enzyme
in
both
rats
and
humans,
with
Km
values
approximately
27­
fold
lower
than
those
for
the
isoenzymes
with
the
next
lowest
value
(
CYP2B1
in
rats,
CYP1A2
in
humans).
The
metabolic
parameters
K
m
and
k
cat
for
rat
and
human
CYP2E1
were
similar.
In
contrast,
the
metabolic
parameters
for
CYP1A2
were
not
similar
across
species.
The
study
authors
concluded
that
the
results
of
this
study
appear
consistent
with
observations
in
vivo
for
the
rat
(
Allis
et
al.,
2002)
and
with
predictions
of
the
existing
PBPK
model
for
bromodichloromethane
in
the
rat
(
Lilly
et
al.,
1998).

Zhao
and
Allis
(
2002)
determined
kinetic
constants
for
metabolism
of
bromodichloromethane
by
CYP2E1,
CYP1A2,
and
CYP3A4
in
human
liver
microsomes.
Constants
for
individual
isoenzymes
were
determined
by
addition
of
enzyme­
specific
inhibitory
antibodies
for
two
isoenzymes
to
the
microsomal
preparations
while
measuring
the
activity
of
the
third.
Measurements
were
performed
in
microsomes
obtained
from
four
donors.
CYP2E1
was
found
to
have
the
lowest
K
m
(
2.9
µ
M)
and
the
highest
catalytic
activity.
The
K
m
values
for
CYP1A2
and
CYP3A4
were
approximately
20­
fold
higher
(
60
µ
M)
and
the
catalytic
activity
was
lower.
Eleven
additional
human
microsomal
preparations
were
characterized
for
activity
of
10
Draft
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February
20,
2003
III
­
13
CYP
isoenzymes.
The
initial
rate
of
metabolism
in
each
preparation
measured
at
9.7
µ
M
bromodichloromethane
was
compared
to
the
activity
of
individual
isoenzymes.
Statistical
analysis
showed
a
significant
correlation
only
with
CYP2E1
activity
at
the
tested
concentration.
The
results
of
this
suggest
that
CYP2E1
would
dominate
metabolism
at
the
low
level
of
exposure
expected
from
ingestion
of
drinking
water.
,
study
However,
the
study
authors
have
noted
that
humans
are
highly
variable
in
the
induction
of
CYP
isoenzymes
and
that
the
contributions
of
the
three
isoenzymes
to
metabolism
of
bromodichloromethane
in
individuals
may
not
be
entirely
predictable.

D.
Excretion
Mink
et
al.
(
1986)
compared
the
excretion
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
male
Sprague­
Dawley
rats
and
male
B6C3F
1
mice.
Animals
were
given
single
oral
doses
of
14C­
labeled
compound
in
corn
oil
by
gavage
at
dose
levels
of
100
mg/
kg
and
150
mg/
kg
for
rats
and
mice,
respectively.
The
lung
was
the
principal
route
of
excretion
in
both
species,
accounting
for
45%
to
88%
of
the
administered
label,
either
as
carbon
dioxide
or
as
parent
compound.
Small
amounts
of
label
(
1.1%
to
4.9%)
were
recovered
in
urine,
but
the
chemical
identity
of
labeled
compounds
was
not
investigated.

Mathews
et
al.
(
1990)
exposed
Fischer
344
rats
to
either
a
single
oral
dose
of
1,
10,
32,
or
100
mg/
kg,
or
10­
day
repeated
doses
of
10
or
100
mg/
kg­
day
bromodichloromethane
dissolved
in
corn
oil.
Approximately
70%
to
80%
of
the
administered
dose
was
excreted
in
exhaled
air
as
14Ccarbon
dioxide,
with
3%
to
5%
as
14C­
carbon
monoxide.
In
general,
less
than
5%
of
the
dose
was
excreted
in
the
urine
or
feces.

E.
Bioaccumulation
and
Retention
No
data
were
located
regarding
the
bioaccumulation
or
retention
of
brominated
trihalomethanes
following
repeated
exposures.
However,
based
on
the
rapid
excretion
and
metabolism
of
the
brominated
trihalomethanes
and
the
low
levels
of
the
structurally­
related
compound
chloroform
detected
in
human
post­
mortem
tissue
samples,
marked
accumulation
and
retention
of
these
compounds
are
not
anticipated.

F.
Summary
No
data
on
absorption
of
brominated
trihalomethanes
were
available
for
humans.
Measurements
in
mice
and
rats
indicate
that
gastrointestinal
absorption
of
brominated
trihalomethanes
is
rapid
(
peak
levels
attained
less
than
an
hour
after
administration
of
a
gavage
dose)
and
extensive
(
63%
to
93%).
Most
studies
of
brominated
trihalomethane
absorption
have
used
oil­
based
vehicles.
A
study
in
rats
found
that
the
initial
absorption
rate
of
bromodichloromethane
was
higher
when
the
compound
was
administered
in
an
aqueous
vehicle
when
compared
to
administration
in
a
corn
oil
vehicle.

Data
for
distribution
of
brominated
trihalomethanes
in
human
organs
and
tissues
are
Draft
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February
20,
2003
III
­
14
limited.
Bromoform
was
found
primarily
in
the
stomach
and
lungs
of
a
human
overdose
victim,
with
lower
levels
detected
in
intestine,
liver,
kidney
and
brain.
Dibromochloromethane
was
found
in
1
of
42
samples
of
human
breast
milk
collected
from
women
living
in
urban
areas.
Radiolabeled
brominated
trihalomethanes
were
detected
in
a
variety
of
tissues
following
oral
dosing
in
rats
and
mice.
Approximately
1
to
4%
of
the
administered
dose
was
recovered
in
body
tissues
when
analysis
was
conducted
8
or
24
hours
post­
treatment.
The
highest
concentrations
were
detected
in
stomach,
liver,
blood,
and
kidneys
when
assayed
8
hours
after
administration
of
the
compounds.
Analyses
of
placentas,
amniotic
fluid
and
fetuses
from
female
rats
and
rabbits
administered
bromodichloromethane
in
drinking
water
indicate
that
this
compound
does
not
accumulate
in
these
tissues
or
fluids.
There
are
no
data
which
are
suggestive
of
strain
specific
differences
in
metabolism.

Brominated
trihalomethanes
are
extensively
metabolized
by
animals.
Metabolism
of
brominated
trihalomethanes
occurs
via
two
pathways.
One
pathway
predominates
in
the
presence
of
oxygen
(
the
oxidative
pathway)
and
the
other
predominates
under
conditions
of
low
oxygen
tension
(
the
reductive
pathway).
In
the
presence
of
oxygen,
the
initial
reaction
product
is
trihalomethanol
(
CX
3
OH),
which
spontaneously
decomposes
to
yield
the
corresponding
dihalocarbonyl
(
CX
2
O).
The
dihalocarbonyl
species
are
quite
reactive
and
may
form
adducts
with
cellular
molecules.
When
intracellular
oxygen
levels
are
low,
the
trihalomethane
is
metabolized
via
the
reductive
pathway,
resulting
in
a
highly
reactive
dihalomethyl
radical
(°
CHX
2),
which
may
also
form
covalent
adducts
with
cellular
molecules.
The
metabolism
of
brominated
trihalomethanes
and
chloroform
appear
to
occur
via
the
same
pathways,
although
in
vitro
and
in
vivo
data
suggest
that
metabolism
via
the
reductive
pathway
occurs
more
readily
for
brominated
trihalomethanes.
Both
oxidative
metabolism
and
reductive
metabolism
of
trihalomethanes
appear
to
be
mediated
by
cytochrome
P450
isoforms.
The
identity
of
cytochrome
P450
isoforms
that
metabolize
brominated
trihalomethanes
has
been
investigated
in
several
studies
which
used
bromodichloromethane
as
a
substrate.
The
available
data
suggest
that
the
cytochrome
P450
isoforms
CYP2E1,
CYP2B1/
2,
and
CYP1A2
metabolize
bromodichloromethane
in
rats.
The
human
isoforms
CYP2E1,
CYP1A2,
and
CYP3A4
show
substantial
activity
toward
bromodichloromethane
in
vitro
and
low
but
measurable
levels
of
CYP2A6
activity
have
also
been
detected.
Based
on
the
available
data,
CYP2E1
and
CYP1A2
are
the
only
isoforms
active
in
both
rats
and
humans.
CYP2E1
shows
the
highest
affinity
for
bromodichloromethane
in
both
species
and
the
metabolic
parameters
K
m
and
k
cat
are
similar
for
rat
and
human
CYP2E1.
In
contrast,
the
metabolic
parameters
for
CYP1A2
differ
in
rats
and
humans.
The
pattern
of
results
for
isozyme
activity
obtained
from
an
inhalation
study
of
bromodichloromethane
was
similar
to
the
pattern
reported
for
male
F344
rats
treated
with
bromodichloromethane
by
gavage.

The
lung
is
the
principle
route
of
excretion
in
rats
and
mice.
Studies
with
14C­
labeled
compounds
indicate
that
up
to
88%
of
the
administered
dose
can
be
found
in
exhaled
air
as
carbon
dioxide,
carbon
monoxide,
and
parent
compound.
Excretion
in
the
urine
generally
appears
to
be
5%
or
less
of
the
administered
oral
dose.
Data
from
one
study
suggests
that
fecal
excretion
is
less
than
3%
of
the
administered
dose.
Draft
­
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20,
2003
III
­
15
Draft
­
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2003
IV
­
1
IV.
HUMAN
EXPOSURE
A.
Occurrence
in
Drinking
Water
The
occurrence
of
brominated
trihalomethanes
in
U.
S.
drinking
water
has
been
determined
in
both
national­
scale
and
localized
studies.
The
occurrence
of
bromodichloromethane
and
bromoform
has
been
described
in
eleven
national
surveys.
Dibromochloromethane
occurrence
has
been
described
in
twelve
national
surveys.
Nine
localized
studies
on
the
occurrence
of
brominated
trihalomethanes
are
also
described
below.

It
is
important
to
note
that
a
variety
of
sampling
and
preservation
techniques
are
used
for
collection
of
occurrence
data
on
brominated
trihalomethanes.
The
addition
of
chlorine
to
raw
water
as
a
disinfectant
at
water
treatment
plants
results
in
the
formation
of
hypochlorous
acid
in
the
processed
water.
The
acid
in
turn
reacts
with
organic
materials
to
produce
chloroform
and
also
oxidizes
available
bromide
ions
to
form
hypobromous
acid.
Hypobromous
acid
reacts
with
organic
materials
in
the
processed
water
to
form
the
brominated
trihalomethanes.
Because
these
chemical
reactions
occur
over
periods
of
days
in
treated
waters,
the
method
used
to
sample
drinking
waters
can
affect
the
measured
concentrations
of
trihalomethanes
in
the
water.
Therefore,
appropriate
sampling
and
preservation
methods
must
be
selected
to
ensure
that
the
analytical
data
are
representative
of
the
desired
endpoint.
For
example,
if
an
investigator
wants
to
know
the
concentration
of
trihalomethanes
in
the
water
at
the
time
of
sampling,
a
reducing
agent
is
added
to
the
sample
containers
to
"
quench"
or
prevent
further
formation
of
trihalomethanes.
If
an
investigator
wants
to
know
the
maximum
amount
of
trihalomethanes
that
could
occur,
no
quenching
is
used
and
the
reactions
are
allowed
to
run
to
completion
at
room
temperature.
If
a
concentration
similar
to
that
at
a
household
tap
is
desired
(
i.
e.,
after
the
water
spends
several
days
in
the
distribution
system,
the
samples
generally
are
not
quenched
but
are
refrigerated
to
slow
the
reactions
(
Wallace,
1997).
Information
on
sample
handling
has
been
included
in
the
discussion
of
individual
studies
when
available
in
the
materials
reviewed
for
this
document.

Spatial
and
temporal
variability
exist
in
the
occurrence
data
reported
for
brominated
trihalomethanes.
Multiple
factors
contribute
to
this
variability.
With
respect
to
spatial
variability,
the
geographical
distribution
of
bromide
ion
in
soil
is
not
uniform
(
Shacklette
and
Boerngen,
1984).
Brominated
byproducts
may
predominate
or
comprise
a
substantial
proportion
of
the
disinfection
byproduct
profile
in
regions
with
high
soil
concentrations.
Brominated
trihalomethanes
may
continue
to
form
within
water
distribution
systems
due
to
the
action
of
free
residual
chlorine
on
remaining
humic
precursors,
resulting
in
substantial
intra­
system
spatial
variability
(
Chen
and
Weisel,
1999).
Temporal
variability
may
result
from
seasonal
variation
in
the
concentration
of
brominated
trihalomethanes
as
a
result
of
seasonal
fluctuations
in
precursor
material
(
Brett
et
al.,
1979).
Short
term
variability
may
be
introduced
by
changes
in
the
demand
cycle
to
individual
homes
or
neighborhoods.
Draft
­
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or
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February
20,
2003
IV
­
2
1.
National
Surveys
The
National
Organics
Reconnaissance
Survey
(
NORS),
conducted
by
U.
S.
EPA,
collected
drinking
water
samples
from
80
cities
nationwide
in
1975.
The
survey
sampled
for
several
organics,
including
brominated
trihalomethanes,
at
the
water
treatment
facilities.
The
sampling
method
employed
was
refrigeration
without
quenching;
therefore,
brominated
trihalomethane
concentrations
may
have
increased
following
collection.
Bromodichloromethane
was
found
in
98%
of
the
systems
sampled.
The
median
concentration
was
8
µ
g/
L
(
ppb),
and
the
maximum
level
was
116
µ
g/
L
(
ppb).
Dibromochloromethane
was
found
in
90%
of
the
systems
sampled
at
a
median
concentration
of
2
µ
g/
L
(
ppb).
The
detection
limit
for
dibromochloromethane
and
bromodichloromethane
was
0.1
µ
g/
L
(
ppb).
The
median
concentration
for
bromoform
was
below
the
detection
limit
of
approximately
5
µ
g/
L
(
ppb)
(
Symons
et
al,.
1975).
NORS
was
performed
prior
to
the
promulgation
of
the
total
trihalomethane
regulation;
therefore,
these
results
may
be
higher
than
current
levels.

The
National
Organics
Monitoring
Survey
(
NOMS)
was
conducted
by
the
EPA
from
March
1976
to
January
1977
(
Wallace,
1997).
In
NOMS,
113
community
water
supplies
were
sampled.
Surface
water
was
the
major
source
for
92
of
the
systems,
and
ground
water
was
the
major
source
for
the
remaining
21
systems.
The
NOMS
used
three
sample
storage
methods.
During
Phase
1,
all
samples
were
refrigerated.
In
Phase
2,
the
samples
were
allowed
to
stand
at
20
to
25

C
for
2
to
3
weeks
to
maximize
trihalomethane
formation.
Phase
3
had
two
parts.
The
samples
identified
as
3T
were
allowed
to
stand
an
additional
2
to
3
weeks.
The
samples
identified
as
3Q
were
quenched
by
addition
of
sodium
thiosulfate.
As
expected,
the
highest
trihalomethane
values
occurred
in
Phases
2
and
3T.
Bromodichloromethane
was
detected
in
over
90%
of
the
systems
sampled.
The
median
concentration
under
the
various
sample
storage
conditions
ranged
from
5.9
to
14
µ
g/
L
(
ppb),
and
the
maximum
concentration
was
183
µ
g/
L
(
ppb).
The
mean
concentrations
of
bromodichloromethane
in
Phases
1,
2,
3T,
and
3Q
were
18,
18,
17,
and
9
µ
g/
L
(
ppb),
respectively.
Dibromochloromethane
was
detected
in
73%
of
the
systems
sampled.
The
median
concentration
ranged
from
below
the
detection
limit
to
3
µ
g/
L
(
ppb),
and
the
maximum
value
was
280
µ
g/
L
(
ppb).
The
mean
concentrations
of
dibromochloromethane
in
Phases
1,
2,
3T,
and
3Q
were
8,
12,
11,
and
6
µ
g/
L
(
ppb),
respectively.
The
median
bromoform
concentration
under
all
sampling
conditions
was
below
the
detection
limit
of
0.3
µ
g/
L
(
ppb);
the
maximum
value
was
280
µ
g/
L
(
ppb).
The
mean
concentrations
of
bromoform
Phases
1,
2,
3T,
and
3Q
were
3,
4,
4,
and
2
µ
g/
L
(
ppb),
respectively.
NOMS
was
conducted
before
the
promulgation
of
the
total
trihalomethane
regulation;
therefore,
these
results
may
be
higher
than
current
levels.

The
Community
Water
Supply
Survey
(
CWSS)
was
conducted
by
the
EPA
in
1978.
The
survey
examined
over
1,100
samples,
representing
over
450
water
supply
systems
(
Brass
et
al.,
1981).
The
samples
were
taken
at
the
treatment
plants
and
in
the
distribution
systems.
In
the
CWSS,
94%
of
the
surface
water
supplies
and
33%
of
the
ground
water
supplies
were
positive
for
bromodichloromethane.
For
surface
water
supplies,
the
mean
of
the
positives
and
the
overall
median
were
12
and
6.8
µ
g/
L
(
ppb),
respectively.
The
mean
of
the
positives
for
ground
water
supplies
was
5.8
µ
g/
L
(
ppb),
and
the
overall
median
was
below
the
minimum
reporting
limit
of
0.5
µ
g/
L
(
ppb).
For
dibromochloromethane,
67%
of
the
surface
water
supplies
and
34%
of
the
Draft
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February
20,
2003
IV
­
3
ground
water
supplies
were
positive.
For
surface
water
supplies,
the
mean
of
the
positives
and
the
overall
median
were
5.0
and
1.5
µ
g/
L
(
ppb),
respectively.
The
mean
of
the
positives
for
ground
water
supplies
was
6.6
µ
g/
L
(
ppb),
and
the
overall
median
was
below
the
minimum
reporting
limit
of
0.5
µ
g/
L
(
ppb).
For
bromoform,
13%
of
the
surface
water
supplies
and
26%
of
the
ground
water
supplies
were
positive.
The
mean
concentration
of
the
positives
in
surface
water
supplies
was
2.1
µ
g/
L
(
ppb),
and
the
overall
median
was
less
than
1.0
µ
g/
L
(
ppb).
The
mean
of
the
positives
for
ground
water
supplies
was
11
µ
g/
L
(
ppb),
and
the
overall
median
was
below
the
minimum
reporting
limit
of
0.5
µ
g/
L
(
ppb)
(
Brass
et
al.,
1981).

The
Rural
Water
Survey
(
RWS)
was
conducted
between
1978
and
1980
by
the
EPA
to
evaluate
the
status
of
drinking
water
in
rural
America.
Samples
from
over
2,000
households,
representing
more
than
600
rural
water
supply
systems,
were
examined.
In
the
RWS,
76%
of
the
surface
water
supplies
and
13%
of
the
ground
water
supplies
were
positive
for
bromodichloromethane
56%
of
the
surface
water
supplies
and
13%
of
the
ground
water
supplies
were
positive
for
dibromochloromethane,
and
18%
of
the
surface
water
supplies
and
12%
of
the
ground
water
supplies
were
positive
for
bromoform.
For
the
surface
water
supplies,
the
mean
of
the
positives
and
the
overall
median
concentrations
were
17
µ
g/
L
(
ppb)
and
11
µ
g/
L
(
ppb)
for
bromodichloromethane,
8.5
µ
g/
L
(
ppb)
and
0.8
µ
g/
L
(
ppb)
for
dibromochloromethane,
and
8.7
µ
g/
L
(
ppb)
and
<
0.5
µ
g/
L
(
ppb)
for
bromoform.
For
the
ground
water
supplies,
the
mean
of
the
positives
was
7.7
µ
g/
L
(
ppb)
for
bromodichloromethane,
9.9
µ
g/
L
(
ppb)
for
dibromochloromethane
and
12
µ
g/
L
(
ppb)
for
bromoform.
The
overall
median
for
ground
water
supplies
was
below
the
minimum
reporting
limit
of
0.5
µ
g/
L
(
ppb)
for
all
three
brominated
trihalomethanes
(
Brass,
1981).

The
Ground
Water
Supply
Survey
(
GWSS)
was
conducted
from
December
1980
to
December
1981
by
the
EPA
to
develop
data
on
the
occurrence
of
volatile
organic
chemicals
in
ground
water
supplies.
Out
of
a
total
of
945
ground
water
systems
that
were
sampled,
466
systems
were
chosen
at
random,
and
the
remaining
479
systems
were
chosen
on
the
basis
of
location
near
industrial,
commercial,
and
waste
disposal
activities.
Samples
were
collected
at
or
near
the
entry
to
the
distribution
system,
and
trihalomethane
formation
was
allowed
to
continue
without
quenching
after
sample
collection.
For
bromodichloromethane,
the
median
of
the
positives
for
the
randomly
chosen
systems
serving
greater
than
10,000
people
was
1.4
µ
g/
L
(
ppb),
and
the
occurrence
rate
was
36%.
For
the
randomly
chosen
smaller
systems,
the
median
positive
concentration
was
1.6
µ
g/
L
(
ppb),
and
the
occurrence
rate
was
54%.
The
nonrandomly
chosen
systems
had
a
median
positive
concentration
of
2.1
µ
g/
L
(
ppb)
and
an
occurrence
rate
of
51%.
For
dibromochloromethane,
the
median
positive
concentration
and
the
occurrence
rate
for
the
randomly
chosen
systems
serving
greater
than
10,000
people
were
2.1
µ
g/
L
(
ppb)
and
31%,
respectively;
these
values
for
the
smaller
systems
were
2.9
µ
g/
L
(
ppb)
and
52%.
The
nonrandomly
chosen
systems
had
a
median
positive
concentration
of
3.9
µ
g/
L
(
ppb)
and
an
occurrence
rate
of
46%.
For
bromoform,
the
median
positive
concentration
was
2.4
µ
g/
L
(
ppb)
for
the
randomly
chosen
systems
serving
greater
than
10,000
and
3.8
µ
g/
L
(
ppb)
for
the
randomly
chosen
systems
serving
fewer
than
10,000
people,
with
occurrence
rates
of
16%
and
31%,
respectively.
The
nonrandomly
chosen
systems
had
a
median
positive
concentration
of
4.2
µ
g/
L
(
ppb)
and
an
occurrence
rate
of
31%
(
Westrick
et
al.,
1983).
Draft
­
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February
20,
2003
IV
­
4
The
National
Screening
Program
for
Organics
in
Drinking
Water
(
NSP),
sponsored
by
the
EPA,
was
conducted
from
June
1977
to
March
1981
and
sampled
169
systems
nationwide.
Samples
were
collected
at
the
treatment
facilities.
For
dibromochloromethane,
the
mean
and
median
for
130
positives
were
17.2
and
10
µ
g/
L
(
ppb),
respectively.
The
maximum
concentration
found
was
131
µ
g/
L
(
ppb)
(
Boland,
1981).

The
Technical
Support
Center
(
TSC)
of
the
Office
of
Ground
Water
and
Drinking
Water
(
OGWDW)
maintains
a
ground
water
contaminant
database.
For
both
bromodichloromethane
and
dibromochloromethane,
the
database
contains
4,439
samples
taken
at
the
treatment
facilities
from
nineteen
states
between
1984
and
1991.
For
bromodichloromethane,
the
mean
concentration
was
5.6
µ
g/
L
(
ppb),
and
the
median
was
3
µ
g/
L
(
ppb).
For
dibromochloromethane,
the
mean
concentration
was
3.0
µ
g/
L
(
ppb),
and
the
median
was
1.7
µ
g/
L
(
ppb).
For
bromoform,
the
database
contains
1409
samples
from
19
states
taken
at
treatment
facilities
between
1984
and
1991.
The
mean
and
median
concentrations
were
determined
to
be
2.5
µ
g/
L
(
ppb)
and
1
µ
g/
L
(
ppb),
respectively
(
U.
S.
EPA,
1991).

Thirty­
five
water
utilities
nationwide,
10
of
which
were
located
in
California,
were
sampled
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
the
clearwell
effluent.
Samples
were
taken
for
four
quarters
(
spring,
summer,
and
fall
in
1988
and
winter
in
1989).
The
median
bromodichloromethane
concentration
for
all
four
quarters
was
6.6
µ
g/
L
(
ppb),
with
the
medians
of
the
individual
quarters
reported
as
6.9,
10,
5.5
and
4.1
µ
g/
L
(
ppb),
respectively,
and
with
a
maximum
value
of
82
µ
g/
L
(
ppb).
For
all
four
quarters,
75%
of
the
measured
concentrations
were
less
than
14
µ
g/
L
(
ppb).
The
median
dibromochloromethane
concentration
for
all
four
quarters
was
3.6
µ
g/
L
(
ppb),
with
the
medians
of
the
individual
quarters
reported
as
2.6,
4.5,
3.8
and
2.7
µ
g/
L
(
ppb),
respectively,
and
with
a
maximum
value
of
63
µ
g/
L
(
ppb).
For
all
four
quarters,
75%
of
the
data
were
below
9.1
µ
g/
L
(
ppb).
The
median
bromoform
concentration
for
all
four
quarters
was
0.57
µ
g/
L
(
ppb),
with
the
medians
of
the
individual
quarters
reported
as
0.33,
0.57,
0.88,
and
0.51
µ
g/
L
(
ppb),
respectively,
and
with
a
maximum
value
of
72
µ
g/
L
(
ppb).
For
all
four
quarters,
75%
of
the
bromoform
concentrations
were
below
2.8
µ
g/
L
(
ppb)
(
Krasner
et
al.,
1989;
U.
S.
EPA
1989a;
1989b).

The
EPA's
Technical
Support
Center
compiled
a
database
from
its
disinfection
by­
products
field
studies.
The
studies
included
a
chlorination
by­
products
survey,
conducted
from
October
1987
to
March
1989.
In
this
survey,
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
determined
in
finished
water
from
the
treatment
plant
and
in
the
distribution
system.
Systems
using
both
surface
water
sources
and
ground
water
sources
were
analyzed.

Mean
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
finished
water
at
the
treatment
plants
were
determined
for
surface
water
systems
serving
both
greater
than
and
less
than
10,000
people.
Forty­
two
samples
were
taken
from
systems
serving
more
than
10,000
people,
and
20
samples
were
taken
from
systems
serving
fewer
than
10,000
people.
The
mean
concentration
of
bromodichloromethane
was
12.7
µ
g/
L
(
ppb)
in
samples
from
systems
serving
more
than
10,000
people
(
90th
percentile,
25.0
µ
g/
L
(
ppb))
and
17.0
µ
g/
L
(
ppb)
Draft
­
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or
Quote
February
20,
2003
IV
­
5
for
samples
from
the
smaller
systems
(
90th
percentile,
29.5
µ
g/
L
(
ppb)).
The
mean
dibromochloromethane
concentrations
was
4.7
µ
g/
L
(
ppb)
for
samples
from
the
larger
systems
(
90th
percentile,
13.8
µ
g/
L
(
ppb))
and
6.9
µ
g/
L
(
ppb)
for
samples
from
the
smaller
systems
(
90th
percentile,
24.2
µ
g/
L
(
ppb)).
The
mean
concentrations
for
bromoform
were
0.7
µ
g/
L
(
ppb)
(
90th
percentile,
1.5
µ
g/
L
(
ppb))
and
0.9
µ
g/
L
(
ppb)
(
90th
percentile,
4.9
µ
g/
L
(
ppb))
in
samples
from
the
larger
systems
and
samples
from
the
smaller
systems,
respectively
(
U.
S.
EPA,
1992a).

Mean
bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
in
distribution
systems
of
these
surface
water
systems
also
were
analyzed.
Thirty­
nine
samples
were
taken
from
systems
serving
more
than
10,000
people,
and
11
samples
were
from
systems
serving
fewer
than
10,000
people.
The
mean
bromodichloromethane
concentrations
in
the
larger
systems
and
the
smaller
systems
were
17.4
µ
g/
L
(
ppb)
(
90th
percentile,
35.3
µ
g/
L
(
ppb))
and
24.8
µ
g/
L
(
ppb)
(
90th
percentile,
51.0
µ
g/
L
(
ppb)),
respectively.
The
mean
dibromochloromethane
concentrations
were
6.3
µ
g/
L
(
ppb)
(
90th
percentile,
17.3
µ
g/
L
(
ppb))
and
10.4
µ
g/
L
(
ppb)
(
90th
percentile,
35.0
µ
g/
L
(
ppb)),
respectively.
Mean
bromoform
concentrations
were
0.8
µ
g/
L
(
ppb)
(
90th
percentile,
3.1
µ
g/
L
(
ppb))
and
1.4
µ
g/
L
(
ppb)
(
90th
percentile,
5.1
µ
g/
L
(
ppb)),
respectively
(
U.
S.
EPA,
1992a).

Ground
water
systems
serving
less
than
10,000
people
were
analyzed
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
seven
finished
water
samples
and
in
five
distribution
system
samples.
Mean
bromodichloromethane
concentrations
in
the
finished
water
samples
and
in
the
distribution
system
samples
were
1.1
µ
g/
L
(
ppb)
(
90th
percentile,
2.6
µ
g/
L
(
ppb))
and
2.2
µ
g/
L
(
ppb)
(
90th
percentile,
5.4
µ
g/
L
(
ppb)),
respectively.
Mean
dibromochloromethane
concentrations
were
0.6
µ
g/
L
(
ppb)
(
90th
percentile,
1.0
µ
g/
L
(
ppb))
and
1.8
µ
g/
L
(
ppb)
(
90th
percentile,
3.6
µ
g/
L
(
ppb)),
respectively.
Mean
bromoform
concentrations
were
0.6
µ
g/
L
(
ppb)
(
90th
percentile,
2.6
µ
g/
L
(
ppb))
and
2.3
µ
g/
L
(
ppb)
(
90th
percentile,
10
µ
g/
L
(
ppb)),
respectively.

For
ground
water
systems
serving
more
than
10,000
people,
dibromochloromethane
and
bromoform
were
not
detected
in
single
samples
taken
at
the
plant
and
from
the
distribution
system,
based
on
a
detection
limit
of
0.2
µ
g/
L
(
ppb).
Bromodichloromethane
concentrations
in
the
plant
and
distribution
system
samples
were
0.2
and
0.4
µ
g/
L
(
ppb),
respectively
(
U.
S.
EPA,
1992a).

The
U.
S.
Geological
Survey
conducted
an
assessment
of
volatile
organic
compounds
in
untreated
ambient
groundwater
of
the
conterminous
United
States
based
on
samples
collected
between
1985
and
1995
from
2948
wells.
The
sampled
wells
were
located
in
rural
and
urban
areas
and
included
wells
used
for
drinking
and
non­
drinking
water
purposes.
A
minimum
reporting
level
of
0.2

g/
L
(
ppb)
was
used
for
most
of
the
compounds,
including
bromodichloromethane,
dibromochloromethane,
and
bromoform.
In
samples
from
the
406
urban
wells
assessed,
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
detected
in
3.0%,
2.8%,
and
2.8%
of
the
wells
examined,
respectively.
In
samples
from
the
2542
rural
wells
examined,
these
compounds
were
detected
in
0.8%,
0.6%,
and
0.4%
of
the
wells,
respectively.
The
measured
concentration
of
the
compounds
in
well
water
were
reported
in
summary
graphics
Draft
­
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or
Quote
February
20,
2003
IV
­
6
only.
Thus,
the
values
reported
here
are
approximate
based
on
visual
inspection
of
the
figures.
The
median
concentrations
measured
in
the
positive
samples
from
the
urban
wells
was
approximately
1.0

g/
L
(
ppb)
for
all
three
compounds,
while
the
maximum
concentrations
of
bromodichloro­
methane,
dibromochloromethane,
and
bromoform
in
the
urban
wells
were
approximately
11,
11,
and
13

g/
L
(
ppb),
respectively.
The
median
concentrations
measured
in
the
positive
samples
from
the
rural
wells
were
approximately
0.4
to
0.5

g/
L
(
ppb)
for
all
three
compounds,
while
the
maximum
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
the
rural
wells
were
approximately
7,
10,
and
18

g/
L
(
ppb),
respectively.

The
most
recent
survey
of
the
occurrence
of
brominated
trihalomethanes
in
public
water
supplies
(
PWSs)
serving
at
least
100,000
persons
resulted
from
the
Information
Collection
Rule
(
ICR)
promulgated
in
May
of
1996
for
disinfectants
and
disinfection
byproducts
(
D/
DBPs).
The
rule
covered
both
surface
and
ground
water
systems.
Monitoring
data
were
collected
from
about
300
water
systems
operating
501
plants
over
the18­
month
period
between
July
1997
and
December
1998.
At
each
plant,
samples
were
collected
monthly
and
analyzed
for
a
variety
of
D/
DBPs
on
a
monthly
or
quarterly
basis.
Bromodichloromethane,
dibromochloromethane,
and
bromoform
were
among
the
analytes
evaluated
quarterly
(
U.
S.
EPA,
2001a).
Five
samples
were
taken
each
quarter
at
each
plant
 
one
of
the
finished
water
and
four
of
the
water
in
the
distribution
system.
Of
the
four
samples
from
the
distribution
system,
one
represented
a
sample
with
the
same
residence
time
as
a
finished
water
sample
held
for
a
specific
period
of
time,
two
represented
approximate
average
water
residence
times
in
the
system,
and
one
sample
was
taken
where
water
residence
time
in
the
system
is
the
longest.
For
each
plant
and
reporting
period,
EPA
compiled
several
summary
statistics.
The
Distribution
System
(
DS)
Average
value
is
the
average
of
the
four
distribution
system
samples.
The
DS
High
Value
is
the
highest
concentration
of
the
four
distribution
system
samples
collected
by
a
plant
in
a
given
quarter.
The
DS
High
Value
might
be
from
any
of
the
four
samples
and
could
vary
from
quarter
to
quarter
depending
on
which
sample
yielded
the
highest
concentrations
in
each
quarter
(
U.
S.
EPA,
2001a).
Table
IV­
1
summarizes
the
results
of
all
six
of
the
quarterly
reporting
periods.

U.
S.
EPA
set
a
minimum
reporting
level
(
MRL)
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
of
1.0

g/
L
for
the
ICR.
The
MRL
is
a
level
below
which
systems
were
not
required
to
report
their
monitoring
results,
even
if
there
were
detectable
results.
Values
below
the
MRL
were
assigned
a
value
of
zero
for
the
purpose
of
calculating
averages;
this
assignment
affects
the
calculation
of
mean
values
for
finished
water
and
DS
high
results
and
calculation
of
all
DS
average
values.

Recent
data
for
concentrations
of
brominated
trihalomethanes
are
now
available
for
117
small
surface
water
plants
(
serving

10,000
people)
from
the
National
Rural
Water
Association
Survey
(
NWRA)
(
U.
S.
EPA
2001b).
Most,
but
not
all,
plants
that
participated
in
the
survey
took
two
samples
at
each
of
three
sampling
locations.
One
sample
was
taken
between
November,
1999,
and
March,
2000,
and
the
other
between
July
and
November,
2000,
for
a
total
of
217
THM
samples.
The
samples
were
taken
at
the
finished
water
location,
distribution
system
average
Draft
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February
20,
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IV
­
7
residence
time
location,
and
maximum
residence
time
location.
These
data
are
summarized
in
Table
IV­
2
below.
Draft
­
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or
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February
20,
2003
IV
­
8
Table
IV­
1.
Brominated
Trihalomethane
Concentrations
Measured
in
U.
S.
Public
Drinking
Water
Systems
Serving
100,000
or
More
Persons
Source
Data
Type
(
a)
Number
of
Samples
Median
(
b)
Mean
(
b)
90th
Percentile
Range
Bromodichloromethane
(

g/
L)

Suface
Water
Finished
1856
6.6
8.2
17.5
<
1.0
­
49
DS
Average
1656
8.6
10.2
20.3
0
­
65.8
DS
High
1656
9.9
11.9
23.3
<
1.0
­
73
Ground
Water
Finished
604
<
1.0
7.9
6.80
<
1.0
­
27
DS
Average
603
1.80
4.06
11.2
0
­
35.3
DS
High
603
2.8
5.78
16.0
<
1.0
­
110
Dibromochloromethane
(

g/
L)

Surface
Water
Finished
1853
1.9
4.03
12.0
<
1.0
­
55.1
DS
Average
1655
2.40
4.72
13.2
0
­
67.3
DS
High
1655
2.9
5.57
15.0
<
1.0
­
67.3
Ground
Water
Finished
604
<
1.0
1.38
4.10
<
1.0
­
33
DS
Average
602
1.35
3.09
8.94
0
­
37.5
DS
High
602
2.1
4.60
12.9
<
1.0
­
85
Bromoform
(

g/
L)

Surface
Water
Finished
1853
<
1.0
0.998
2.88
<
1.0
­
34
DS
Average
1653
0
1.18
3.10
0
­
34.3
DS
High
1653
<
1.0
1.48
3.90
<
1.0
­
40
Ground
Water
Finished
602
<
1.0
0.838
2.20
<
1.0
­
21
DS
Average
599
0.325
1.92
4.78
0
­
28.8
DS
High
599
1.2
2.95
7.72
<
1.0
­
391
(
a)
Finished
=
sample
location
after
treatment,
before
entering
the
distribution
system
(
DS);
DS
Average
=
average
of
four
sample
locations
in
the
DS;
DS
High
=
the
highest
concentration
of
the
four
distribution
system
samples
collected
by
a
plant
in
a
given
quarter.
For
purposes
of
calculations,
all
values
below
the
minimum
reporting
level
(
MRL)
of
1.0

g/
L
for
all
three
compounds
were
assigned
a
value
of
zero.
(
b)
Median
and
mean
of
all
samples
including
those
below
the
MRL.
Source:
Disinfectants
and
Disinfection
Byproducts
(
D/
DBPs)
ICR
Data,
U.
S.
EPA
(
2001a).
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
9
Table
IV­
2
NRWA
Brominated
Trihalomethane
Results
for
Small
Surface
Water
Plants
THM
Sample
Location
Mean
Median
90th
Percentile
Range
Bromodichloromethane
(

g/
L)
Finished
11.2
6.5
26.6
0
­
84.4
DS
Average
14.3
9.4
32.2
0
­
100.3
DS
Max
15.9
10.2
34.2
0
­
121.1
Dibromochloromethane
(

g/
L)
Finished
5.0
1.1
13.4
0
­
83.1
DS
Average
6.1
1.5
16.3
0
­
99.0
DS
Max
6.7
1.9
17.1
0
­
91.6
Bromoform
(

g/
L)
Finished
4.0
0
1.2
0
­
333.4
DS
Average
4.6
0
1.2
0
­
340.5
DS
Max
4.5
0
1.3
0
­
349.7
Median
and
mean
of
all
samples,
including
those
below
the
detection
limit.
Source:
National
Rural
Water
Association
Survey.

2.
Other
Studies
Several
less
comprehensive
surveys
have
analyzed
drinking
water
for
one
or
more
of
the
brominated
trihalomethanes.
An
overview
of
these
studies
is
provided
below.

The
EPA
Region
V
Organics
Survey
sampled
finished
water
from
83
sites
in
a
region
that
includes
Illinois,
Indiana,
Michigan,
Minnesota,
Ohio,
and
Wisconsin.
Bromoform
was
found
at
a
median
concentration
of
the
positives
of
1
µ
g/
L
(
ppb)
and
a
maximum
level
of
7
µ
g/
L
(
ppb).
A
total
of
14%
of
the
locations
sampled
contained
detectable
levels
of
bromoform
(
U.
S.
EPA,
1980).
Kelley
(
1985)
surveyed
18
drinking
water
plants
in
Iowa
for
trihalomethanes,
and
detected
bromoform
in
five
water
supplies
at
concentrations
ranging
from
1.0
to
10
µ
g/
L
(
ppb).

The
EPA's
Five­
year
Total
Exposure
Assessment
Methodology
(
TEAM)
study
measured
the
personal
exposures
of
a
probability­
based
sample
of
residents
in
several
U.
S.
cities
to
various
organic
chemicals
in
air
and
drinking
water
between
1981
and
1987.
As
part
of
the
study,
running
tap
water
samples
were
collected
from
residences
of
nearly
850
study
participants
during
the
morning
and
the
evening
to
test
for
brominated
trihalomethane
concentrations.
The
samples
were
quenched
with
sodium
thiosulfate
at
the
time
of
collection.
Tables
IV­
3,
IV­
4,
and
IV­
5
show
bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
found
in
drinking
water
from
the
six
cities
surveyed.
Samples
of
water
were
taken
from
each
participating
residence
at
the
household
taps
and
sodium
thiosulfate
added
as
a
quenching
agent.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
10
Table
IV­
3
Bromodichloromethane
Concentrations
in
Drinking
Water
from
the
U.
S.
EPA
TEAM
Study
(
µ
g/
L)

Location
Date
Sampled
Sample
Size
%

Measured
Concentration
µ
g/
L
(
ppb)

Mean
Median
Max
Percentiles
25%
75%
90%
95%

Elizabeth/
Bayonne,
New
Jersey
Fall
1981
340
99.7
13.6
13
23
­­
15
16
18
Summer
1982
156
99.8
13.6
12
54
­­
15
18
20
Winter
1983
49
100
5.4
5.8
16
­­
7.1
8.3
8.3
Los
Angeles,
California
Winter
1984
117
93
11
12
23
5.1
16
17
20
Summer
1984
52
96
20
24
38
7.7
31
33
37
Winter
1987
9
89
19
24
31
­­
­­
­­
­­

Summer
1987
7
100
26
27
36
­­
­­
­­
­­

Antioch/
Pittsburg,
California
Spring
1984
71
96
21
17
47
2.4
36
45
47
Devils
Lake,
North
Dakota
Fall
1982
24
73
0.21
0.18
1.0
­­
­­
­­
­­

Greensboro,
North
Dakota
Fall
1982
24
93
7.1
7.8
11
­­
9.2
­­
­­

Baltimore,
Maryland
Spring
1987
10
100
10
10
13
­­
­­
­­
­­

Adopted
from
Hartwell,
(
1987),
Wallace
et
al.
(
1987),
Wallace
et
al.
(
1988),
and
Wallace
(
1992)
by
U.
S.
EPA
(
1994b).
Mean
and
median
values
of
all
samples,
including
those
below
the
quantitation
limit.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
11
Table
IV­
4
Dibromochloromethane
Concentrations
in
Drinking
Water
from
the
U.
S.
EPA
TEAM
Study
Location
Date
Sampled
Sample
Size
%

Measured
Concentration
µ
g/
L
(
ppb)

Mean
Median
Max
Percentile
25%
75%
90%
95%

Elizabeth/
Bayonne,
New
Jersey
Fall
1981
340
99.7
2.4
2.4
8.4
­­
2.7
3.2
3.4
Summer
1982
156
99.8
2.1
1.9
7.2
­­
2.4
3.1
3.8
Winter
1983
49
93
1.4
1.6
3.0
­­
1.8
2.0
2.1
Los
Angeles,
California
Winter
1984
117
89
9.4
11
19
2.4
15
17
18
Summer
1984
52
85
28
32
55
15
42
43
48
Winter
1987
9
89
10
12
17
­­
­­
­­
­­

Summer
1987
7
100
24.7
18
70
­­
­­
­­
­­

Antioch/
Pittsburg,
California
Spring
1984
71
85
8
6.4
21
0.98
15
18
19
Devils
Lake,
North
Dakota
Fall
1982
24
18
0.09
0.06
0.45
­­
0.06
­­
­­

Greensboro,
North
Dakota
Fall
1982
24
93
1.2
1.2
1.9
­­
1.5
­­
­­

Baltimore,
Maryland
Spring
1987
10
100
2.7
2.6
3.5
­­
­­
­­
­­

Adopted
from
Hartwell,
(
1987),
Wallace
et
al.
(
1987),
Wallace
et
al.
(
1988),
and
Wallace
(
1992)
by
U.
S.
EPA
(
1994b).
Mean
and
median
values
of
all
samples,
including
those
below
the
quantitation
limit.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
12
Table
IV­
5
Bromoform
Concentrations
in
Drinking
Water
from
the
U.
S.
EPA
TEAM
Study
Location
Date
Sampled
Sample
Size
%

Measured
Concentration
µ
g/
L
(
ppb)

Mean
Median
Max
Percentile
25%
75%
90%
95%

Los
Angeles,
California
Winter
1984
117
69
0.78
0.54
12
0.34
0.92
1.2
1.5
Summer
1984
52
90
8.08
3.0
78
2.0
5.9
13
53
Winter
1987
9
89
3.2
3.2
4.7
­­
­­
­­
­­

Summer
1987
7
100
25.5
9.6
113
­­
­­
­­
­­

Antioch/
Pittsburg,
California
Spring
1984
71
69
0.78
0.58
2.0
0.19
1.2
1.8
1.9
Adopted
from
Wallace
(
1992)
by
U.
S.
EPA
(
1994b).
Mean
and
median
values
of
all
samples,
including
those
below
the
quantitation
limit.
Bromoform
was
measured
in
fewer
than
10%
of
samples
from
the
other
four
cities
in
the
TEAM
study
and
are
not
presented
here
Furlong
and
D'itri
(
1986)
reported
that
a
survey
of
water
treatment
plants
in
Michigan
detected
bromodichloromethane
in
35
of
40
plants
at
a
median
concentration
of
2.7
µ
g/
L
(
ppb)
and
a
maximum
of
54.2
µ
g/
L
(
ppb);
the
mean
of
the
positive
samples
was
7.4
µ
g/
L
(
ppb).

Dibromochloromethane
was
also
detected
in
30
plants
at
a
median
concentration
of
2.2
µ
g/
L
(
ppb)
and
a
maximum
of
39.6
µ
g/
L
(
ppb);
the
mean
of
the
positives
was
5.1
µ
g/
L
(
ppb).
Bromoform
was
detected
at
three
of
40
plants
sampled
at
concentrations
of
0.9,
1.3,
and
1.6
µ
g/
L
(
ppb).

Fair
et
al.
(
1988)
analyzed
drinking
water
from
three
community
water
supplies
for
chlorination
by­
products.
Bromodichloromethane
concentrations
ranged
from
7.5
to
30
µ
g/
L
(
ppb)
in
finished
water
and
from
9.9
to
36
µ
g/
L
(
ppb)
in
the
distribution
systems.
Dibromochloromethane
concentrations
ranged
from
less
than
0.5
to
19
µ
g/
L
(
ppb)
in
finished
water
at
the
plant
and
from
less
than
0.5
to
23
µ
g/
L
(
ppb)
in
the
distribution
systems.
Bromoform
concentrations
ranged
from
less
than
0.5
to
2.5
µ
g/
L
(
ppb)
in
finished
water
and
from
less
than
0.5
to
3.1
µ
g/
L
(
ppb)
in
the
distribution
systems.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
13
Wallace
et
al.
(
1982)
analyzed
tap
water
for
bromodichloromethane
as
part
of
a
study
to
determine
individual
exposures
to
volatile
organics
during
normal
daily
activities
of
students
at
the
University
of
North
Carolina,
Chapel
Hill.
Bromodichloromethane
was
detected
in
7
of
7
samples
of
tap
water,
at
concentrations
ranging
from
15
to
20
µ
g/
L
(
ppb),
with
a
mean
of
17
µ
g/
L
(
ppb).
The
detection
limit
was
0.1
µ
g/
L
(
ppb).

Chang
and
Singer
(
1984)
analyzed
the
bromoform
concentration
in
drinking
water
samples
prepared
by
the
desalination
of
seawater.
After
pretreatment
using
either
activated
carbon
or
ultrafiltration,
but
prior
to
reverse
osmosis
treatment,
bromoform
concentrations
were
13
±
14
and
110
±
59
µ
g/
L
(
ppb),
respectively.
After
reverse
osmosis
was
completed,
the
finished
water
product
contained
bromoform
concentrations
ranging
from
2.0
to
51
µ
g/
L
(
ppb)
(
mean,
20.17
µ
g/
L
(
ppb))
when
activated
carbon
was
used
as
a
pretreatment
and
127
µ
g/
L
(
ppb)
when
ultrafiltration
was
used.
In
the
reverse
osmosis
treatment,
three
reverse
osmosis
membranes
were
evaluated.
The
cellulose
triacetate
filter
resulted
in
concentrations
of
51
µ
g/
L
(
ppb),
while
the
polyether/
urea
thin
film
spiral
wound
membrane
and
the
polysulfone
membrane
filters
which
resulted
in
final
concentrations
of
5.0
µ
g/
L
(
ppb)
and
2.25
µ
g/
L
(
ppb),
respectively.

Bromodichloromethane,
dibromochloromethane,
and
bromoform
were
detected
in
9.5
to
12.8%
of
drinking
water
samples
collected
in
1987
in
Nassau
County,
New
York.
The
county
draws
its
drinking
water
from
underground
aquifers.
Bromodichloromethane
and
dibromochloromethane
had
similar
concentration
profiles,
being
detected
in
approximately
10%
and
8.5%
of
the
samples,
respectively,
at
concentrations
less
than
4.9
ppb.
The
detection
limit
was
1
ppb
for
each
chemical.
Bromoform
was
detected
in
8%
of
the
samples
at
2
to
4.9
ppb,
in
2.5%
of
the
samples
at
5
to
9.9
ppb,
and
in
less
than
1%
of
the
samples
at
10
to
49.9
ppb.
The
detection
limit
was
2
ppb.
None
of
the
drinking
water
samples
contained
more
than
50
ppb
of
any
of
the
trihalomethanes,
and
less
than
1%
of
the
samples
contained
between
10
and
49.9
ppb
of
the
brominated
compounds
(
Moon
et
al.
1990).

U.
S.
EPA
conducted
a
study
of
contaminants
in
household
water
in
nine
residences
as
part
of
a
larger
study
of
health
risks
due
to
environmental
contamination
in
the
Lower
Rio
Grande
Valley
(
Berry
et
al.,
1997).
Samples
of
water
used
for
drinking
were
taken
once
during
a
3­
day
period
in
the
spring
and
once
during
a
2­
day
period
in
the
summer
of
1993.
Water
used
for
drinking
in
the
nine
residences
could
be
traced
to
one
of
three
sources:
the
municipal
water
supply
of
Brownsville,
Texas,
vended
water
supplies
(
municipal
water
that
had
undergone
further
treatment),
and
well
water.
Samples
were
collected
using
U.
S.
EPA
protocols,
including
quality
assurance
samples
and
field
blanks.
The
detection
and
minimum
quantitation
limits
for
each
analyte
were
documented
in
other
reports.
Bromodichloromethane,
dibromochloromethane,
and
bromoform
were
detected
in
the
household
water
of
seven
of
the
nine
residences
during
the
spring
and
in
five
of
the
nine
residences
during
the
summer
(
Berry
et
al.,
1997).
During
the
spring,
the
minimum,
median,
and
maximum
concentrations
of
bromodichloromethane
for
the
seven
positive
samples
were
3.2,
5.2,
and
24.4

g/
L
(
ppb),
respectively.
For
dibromochloromethane,
the
values
were
3.3,
5.1,
and
17.3

g/
L
(
ppb),
respectively.
For
bromoform,
the
values
were
1.0,
3.0,
and
14.1

g/
L
(
ppb),
respectively.
During
the
summer,
the
minimum,
median,
and
maximum
concentrations
of
bromodichloromethane
in
the
five
positive
samples
were
2.3,
7.7,
and
34.3

g/
L
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
14
(
ppb),
respectively.
For
dibromochloromethane,
the
values
were
1.8,
7.6,
and
49.9

g/
L
(
ppb),
respectively,
and
for
bromoform,
the
values
were
1.6,
7.8,
and
31.7

g/
L
(
ppb),
respectively.

Weisel
et
al.
(
1999)
examined
concentrations
of
trihalomethanes
in
the
tap
water
of
the
homes
of
49
women
in
New
Jersey.
The
49
residences
were
selected
so
that
approximately
half
would
represent
the
lower
extreme
of
trihalomethane
contamination
and
half
the
upper
extreme
of
trihalomethane
contamination
identified
in
a
previous
study.
Samples
were
stored
unquenched
on
ice
after
collection
and
were
analyzed
within
24
hours.
The
three
brominated
trihalomethanes
were
detected
in
all
49
samples.
The
mean
(
±
standard
deviation)
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
5.7
±
8.6,
2.0
±
2.1,
and
0.73
±
0.90

g/
L
(
ppb),
respectively.
The
median
values
for
the
three
compounds
were
2.6,
1.4,
and
0.45

g/
L
(
ppb),
respectively.
These
values
are
not
representative
of
New
Jersey,
because
of
the
selection
criteria
for
the
residences.
The
ranges
(
minimum
to
maximum)
of
concentrations
measured
for
each
compound
were
0.06
to
48

g/
L
(
ppb)
for
bromodichloromethane,
0.14
to
9.7

g/
L
(
ppb)
for
dibromochloromethane,
and
0.03
to
4.21

g/
L
(
ppb)
for
bromoform.

3.
Estimates
of
Tap
Water
Ingestion
Exposure
to
Brominated
Trihalomethanes
a.
Estimates
Based
on
ICR
Data
for
Disinfection
Byproducts
The
data
from
EPA's
ICR
for
disinfectants
and
disinfection
byproducts
(
U.
S.
EPA
2001a)
offer
several
advantages
over
the
other
national
studies
for
purposes
of
estimating
national
exposure
levels
of
adults
in
the
United
States
to
brominated
trihalomethanes
via
ingestion
of
drinking
water.
First,
they
are
recent
and
reflect
relatively
current
conditions.
Second,
data
of
very
similar
quality
and
quantity
were
collected
systematically
from
a
large
number
of
plants
(
501)
and
systems
(
approximately
300),
including
both
surface
and
ground
water
systems.
Third,
the
mean,
median,
and
90th
percentile
value
were
estimated
on
the
basis
of
all
samples
taken,
not
just
the
sample
detects.
Thus,
these
descriptive
statistics
are
representative
of
the
exposures
of
the
entire
populations
served
by
those
systems,
not
just
the
populations
served
by
systems
with
higher
concentrations
of
these
compounds.
However,
this
study
can
not
be
considered
representative
of
smaller
public
water
supplies
or
water
supplies
from
the
most
highly
industrialized
or
contaminated
areas.

Table
IV­
6
presents
estimated
drinking
water
exposures
to
brominated
trihalomethanes
of
the
adult
populations
served
by
large
public
water
systems
(
serving
100,000
or
more
persons)
based
on
the
ICR
Occurrence
Data
(
U.
S.
EPA,
2001a).
Exposure
was
calculated
by
multiplying
the
concentration
of
individual
brominated
trihalomethanes
in
drinking
water
by
the
average
daily
intake,
assuming
that
each
individual
consumes
two
liters
of
water
per
day.
The
annual
median,
mean,
and
upper
90th
percentile
values
are
presented
for
both
surface
and
ground
water
systems.
Assuming
that
the
DS
High
value
actually
represents
the
average
exposure
level
of
persons
served
by
one
plant
distribution
pipe
with
the
longest
water­
residence
time,
the
DS
High
value
might
be
used
to
estimate
a
high­
end
exposure
level.
Thus,
the
90th
percentile
of
the
DS
High
values
are
also
presented
in
Table
IV­
6.
Draft
­
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February
20,
2003
IV
­
15
Table
IV­
6
Estimated
Drinking
Water
Exposures
to
Brominated
Trihalomethanes
in
U.
S.
Public
Drinking
Water
Systems
Serving
More
than
100,000
Personsa
Source
Medianb
Meanb
90th
Percentileb
DS
High
90th
Percentilec
Bromodichloromethane
(

g/
person/
day)

Surface
Water
17
20
40
47
Ground
Water
3.6
8.1
22
32
Dibromochloromethane
(

g/
person/
day)

Surface
Water
4.8
9.4
26
30
Ground
Water
2.7
6.2
18
26
Bromoform
(

g/
person/
day)

Surface
Water
0
2.4
6.2
7.8
Ground
Water
0.65
3.8
9.6
15
a
Source:
U.
S.
EPA
(
2001a).
Assumes
that
each
individual
consumes
2
liters
of
water
daily.
Also
assumes
that
concentrations
at
the
drinking
water
tap
are
similar
to
concentrations
in
the
distribution
system
(
DS)
sampled
at
locations
considered
to
be
representative
of
average
(
DS
Average)
and
highest
(
DS
High)
retention
times
(
see
Table
IV­
1).
b
Based
on
concentrations
from
the
DS
Average
values.
c
Based
on
the
90th
percentile
of
the
DS
High
values
to
represent
a
plausible
high­
end
exposure
level.

For
bromodichloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
17,
20,
and
40
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromodichloromethane
from
ground
water
systems
are
lower
 
3.6,
8.1,
and
22
µ
g/
person/
day,
respectively.
For
dibromochloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
4.8,
9.4,
and
26
µ
g/
person/
day,
respectively.
The
corresponding
values
for
populations
exposed
to
dibromochloromethane
from
groundwater
system
are
lower
 
2.7,
6.2,
and
18
µ
g/
person/
day,
respectively.
For
bromoform,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
near
0,
2.4,
and
6.2
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromoform
from
ground
water
systems
are
higher
 
0.65,
3.8,
and
9.6
µ
g/
person/
day,
respectively.

Average
daily
intake
of
dibromochloromethane
was
also
evaluated
for
determination
of
the
Relative
Source
Concentration.
The
details
of
this
evaluation
are
presented
in
Appendix
C.
Intake
for
ingestion
was
calculated
using
mean
intake
rates
of
1.2
or
0.6
L/
day
for
total
and
direct
intake
(
NRC,
1999),
respectively.
Direct
intake
includes
consumption
of
water
directly
from
the
tap,
but
does
not
include
intake
of
tap
water
used
for
preparation
of
heated
items
such
tea,
coffee,
or
soup.
Based
on
the
ICR
distribution
system
average
concentration
of
4.72
µ
g/
L
for
Draft
­
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Cite
or
Quote
February
20,
2003
IV
­
16
dibromochloromethane
in
surface
water,
the
average
daily
total
and
direct
and
ingestion
intakes
would
be
5.7
and
2.8
µ
g/
day,
respectively.
Absorption
of
dibromochloromethane
from
tap
water
was
estimated
using
methodology
described
in
U.
S.
EPA
(
1992c),
as
modified
by
Vecchia
and
Bunge
(
2002).
The
average
dermal
uptake
of
dibromochloromethane
was
estimated
to
be
2
µ
g
per
shower
or
bathing
event.
Intake
via
inhalation
of
dibromochloromethane
volatilized
during
household
activities
(
e.
g.,
showering,
bathing,
dishwashing,
toilet
flushing,
etc.)
was
estimated
using
a
three­
compartment
model
based
on
McKone
(
1987).
This
model
estimated
an
average
daily
inhalation
exposure
of
7
µ
g/
day
for
the
volatilized
compound.
Parallel
calculations
were
not
performed
for
bromodichloromethane
or
bromoform,
because
these
compounds
are
probable
carcinogens.
Therefore,
in
accordance
with
U.
S.
EPA
policy,
RSC
analysis
was
not
conducted.

b.
Estimates
of
Ingestion
Exposure
Based
on
Other
National
Studies
Exposure
to
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
drinking
water
from
ground
water
supplies
can
be
estimated
from
the
median
levels
found
in
the
GWSS.
Based
on
the
range
of
median
levels
(
1.4
 
2.1
µ
g/
L
(
ppb))
and
a
consumption
rate
of
two
liters
per
day,
the
median
exposure
to
bromodichloromethane
may
range
from
2.8
to
4.2
µ
g/
day.
Similarly,
median
exposure
to
dibromochloromethane
may
range
from
4.2
to
7.8
µ
g/
day,
and
for
bromoform,
median
exposure
may
range
from
4.8
to
8.4
µ
g/
day.
Exposure
to
bromodichloromethane
from
surface
water
supplies
can
be
estimated
based
on
the
range
of
median
values
observed
under
different
conditions
in
NOMS,
which
mainly
sampled
surface
water
systems.
Based
on
a
range
of
5.9
 
14
µ
g/
L
(
ppb),
exposure
to
bromodichloromethane
from
surface
water
is
estimated
to
be
between
12
and
28
µ
g/
day.
Similarly,
based
on
the
range
of
medians
reported
for
dibromochloromethane
concentrations,
the
median
exposure
is
estimated
to
be
up
to
6
µ
g/
day.
The
median
levels
of
bromoform
in
the
surface
water
supplies
have
been
found
to
be
less
than
the
EPA
Drinking
Water
minimum
reporting
levels
(
MRLs)
of
0.5
 
1
µ
g/
L
(
ppb).
An
estimate
of
exposure
based
on
the
MRLs
will
be
overly
conservative
because
the
actual
concentration
of
bromoform
is
not
detectable.
Based
on
the
range
of
MRLs,
0.5
 
1
µ
g/
L
(
ppb),
the
exposure
to
bromoform
is
estimated
to
range
from
1
to
2
µ
g/
day
for
surface
water
supplies.

Ingestion
exposure
to
brominated
trihalomethanes
in
drinking
water
can
also
be
estimated
from
the
concentrations
found
at
the
tap
in
the
TEAM
studies.
Table
IV­
7
presents
median,
mean,
90th
percentile,
and
95th
percentile
estimates
of
daily
intakes
of
bromodichloromethane,
dibromochloromethane,
and
bromoform,
based
on
an
assumed
drinking
water
ingestion
rate
of
2
liter
per
day.
Table
IV­
7
provides
estimates
for
those
locations
and
seasons
with
a
sample
size
of
at
least
50,
with
one
exception.
Devils
Lake,
ND,
with
a
samples
size
of
only
24,
is
added
to
represent
an
area
with
low
concentrations.
Thus,
the
influence
of
small
sample
size
on
distributional
statistics
should
be
minimized
in
Table
IV­
7.
The
median,
mean,
and
90th
percentile
values
in
Table
IV­
7
for
the
TEAM
study
can
be
compared
with
the
corresponding
values
in
Table
IV­
6
for
the
ICR
Occurrence
data.

Table
IV­
7
demonstrates
that
concentrations
of
brominated
trihalomethanes
are
lower
in
winter
than
in
summer,
as
would
be
expected
on
the
basis
of
temperature.
In
this
sample
of
Draft
­
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February
20,
2003
IV
­
17
Table
IV­
7.
Estimated
Distribution
of
Drinking
Water
Exposures
to
Brominated
Trihalomethanes
for
Populations
in
U.
S.
EPA
TEAM
Study
(
a)

Location
Season
Year
Median
(
b)
Mean
(
b)
90th
Percentile
(
b)
95th
Percentile
Bromodichloromethane
(

g/
person/
day)

Elizabeth/
Bayone
NJ
summer
82
24
27
36
40
winter
83
12
11
17
17
Los
Angeles,
CA
summer
84
48
40
66
74
winter
84
24
22
34
40
Antioch/
Pittsburg,
CA
spring
84
34
42
90
94
Devils
Lake,
ND
fall
82
0.36
0.42
<
2.0
<
2.0
Dibromochloromethane
(

g/
person/
day)

Elizabeth/
Bayone
NJ
summer
82
3.8
4.2
6.2
7.6
winter
83
3.2
2.8
4.0
4.2
Los
Angeles,
CA
summer
84
64
56
86
96
winter
84
22
19
34
36
Antioch/
Pittsburg,
CA
spring
84
13
16
36
38
Devils
Lake,
ND
fall
82
0.12
0.2
<
0.9
<
0.9
Bromoform
(

g/
person/
day)

Los
Angeles,
CA
summer
84
6.0
16.2
26
100
winter
84
1.1
1.6
2.4
3.0
Antioch/
Pittsburg,
CA
spring
84
1.2
1.6
3.6
3.8
(
a)
Intakes
estimated
from
data
in
Tables
IV­
3,
IV­
4,
and
IV­
5
assuming
a
water
ingestion
rate
of
2
liters
per
day.
Selected
locations
and
seasons
with
samples
sizes
over
50.
Added
Devils
Lake,
ND,
to
represent
an
area
with
low
air
concentrations.
(
b)
Median,
mean,
and
upper
percentiles
estimated
for
entire
population
of
city.
Draft
­
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or
Quote
February
20,
2003
IV
­
18
geographic
locations,
estimates
of
the
average
of
the
population
intakes
of
bromodichloromethane
from
drinking
water
range
from
0.42
to
42

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
2.0
to
90

g/
person/
day.
Estimates
of
the
average
population
intake
of
dibromochloromethane
from
drinking
water
range
from
0.2
to
56

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
0.9
to
86

g/
person/
day.
Estimates
of
the
average
of
the
population
intakes
of
bromoform,
for
those
areas
in
which
bromoform
was
measurable
in
a
majority
of
the
samples,
range
from
1.6
to
16.2

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
2.4
to
26

g/
person/
day.
Four
of
the
six
locations
in
the
TEAM
study,
however,
had
a
low
frequency
(
less
than
10%)
of
detection
of
bromoform
in
measurable
quantities.

c.
Sources
of
Uncertainty
in
Estimates
of
Exposure
from
Drinking
Water
Sources
of
uncertainty
in
the
estimates
of
ingestion
exposure
include
use
of
different
analytical
methods,
failure
to
report
quantitation
limits,
using
measures
near
the
detection
limit,
failure
to
report
how
nondetects
are
handled
when
averaging
values
(
e.
g.,
set
to
zero
or
one
half
the
detection
limit),
and
failure
to
report
sample
storage
method
and
duration.
In
addition,
many
environmental
factors
influence
the
concentrations
of
these
compounds
in
drinking
water
at
the
tap
and
in
vended
or
bottled
waters
used
for
drinking.
These
factors
include
season
and
temperature,
geographic
location,
source
of
water,
residence
time
in
distribution
system,
and
others.

B.
Exposure
from
Sources
Other
Than
Drinking
Water
1.
Dietary
Intake
a.
Measured
Concentrations
in
Foods
and
Beverages
Information
on
the
levels
of
brominated
trihalomethanes
in
foods
and
beverages
is
limited.
Chlorine
is
used
in
food
production
for
applications
such
as
the
disinfection
of
chicken
in
poultry
plants
and
the
superchlorination
of
water
at
soda
and
beer
bottling
plants
(
Borum,
1991).
Therefore,
the
possibility
exists
for
contamination
of
foodstuffs
by
disinfection
by­
products
with
resulting
dietary
exposure.
The
occurrence
of
bromodichloromethane
in
foods
and
beverages
is
the
best
characterized
of
the
three
compounds.
Less
information
is
available
concerning
the
occurrence
of
dibromochloromethane
or
bromoform
in
foods
and
beverages
in
the
United
States.
Some
information
is
available
from
international
studies,
but
may
not
be
relevant
to
U.
S.
occurrence
because
of
different
water
treatment
and
food
processing
practices.
The
available
U.
S.
and
international
studies
are
summarized
below.

Entz
et
al.
(
1982)
analyzed
food
samples
from
Elizabeth,
NJ,
Chapel
Hill,
NC,
and
Washington,
DC.
for
bromodichloromethane.
A
total
of
39
different
food
items
from
each
city
were
collected
according
to
standards
set
for
the
FDA's
Total
Diet
Market
Basket
Study.
The
Adult
Market
Basket,
representing
the
diet
of
a
teenage
male,
is
divided
into
12
food
groups.
Individual
foods
are
prepared
as
generally
consumed
in
the
home
and
foods
from
each
group
are
Draft
­
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or
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February
20,
2003
IV
­
19
blended
together
in
"
the
proper
proportions"
to
form
composites.
In
this
study,
foods
were
blended
into
four
composites
representing
dairy
products;
meat,
fish
and
poultry;
oils,
fats
and
shortening;
and
beverages.
The
estimated
limit
of
quantitation
for
bromodichloromethane
in
each
of
these
composites
was
2.3,
4.5,
8.3,
and
0.5
ng/
g,
respectively.
Five
sets
of
each
composite
were
tested
for
a
total
of
20
composites.
Bromodichloromethane
was
detected
in
one
dairy
composite
at
1.2
ppb
and
two
beverage
composites
at
0.3
ppb
and
0.6
ppb.
Analysis
of
individual
foods
from
the
beverage
and
dairy
composites
found
bromodichloromethane
in
three
samples
of
cola
soft
drinks
at
concentrations
of
2.3
ppb,
3.4
ppb,
and
3.8
ppb
and
in
one
sample
of
butter
at
7
ppb.

Uhler
and
Diachenko
(
1987)
sampled
38
food
and
beverage
products
from
15
food
processing
plants
in
nine
states.
Plants
were
chosen
on
a
"
worst­
case"
basis
from
areas
where
contaminated
water
would
most
likely
be
used
in
processing.
In
addition,
processing
plants
were
chosen
for
study
only
if
they
produced
high
fat
content
food
that
came
in
contact
with
water
during
processing
or
contained
a
high
percentage
of
added
water.
Samples
containing
less
than
1
ng/
g
were
considered
nondetects.
Bromodichloromethane
was
detected
in
6
out
of
37
tested
food
tested
at
the
following
levels:
two
samples
of
clear
sodas
at
1.2
and
2.3
ng/
g
(
ppb)
and
one
sample
of
dark
cola
at
l.
2
ng/
g
(
ppb)
out
of
fifteen
soft
drinks,
and
three
of
six
samples
of
ice
cream
at
0.6
to
2.3
ng/
g
(
ppb).
Bromodichloromethane
was
not
found
in
any
of
the
eight
cheese
samples
analyzed.

U.
S.
EPA
(
1985)
reported
that
bromodichloromethane
was
identified
in
bacon.
No
further
information
on
sample
size,
detection
limit,
or
study
methodology
was
provided.

Abdel­
Rahman
(
1982)
analyzed
various
soft
drinks
for
bromodichloromethane
and
found
average
levels
ranging
from
0.2
to
6.6
µ
g/
L
(
ppb)
for
colas
and
from
0.1
to
0.2
µ
g/
L
(
ppb)
for
clear
soft
drinks
(
Abdel­
Rahman,
1982).
In
Italy,
Cocchioni
et
al.
(
1996)
analyzed
61
samples
of
different
commercially
prepared
beverages
and
94
samples
of
mineral
waters
for
volatile
organohalogenated
compounds.
In
the
prepared
beverages,
they
found
maximum
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
of
40.6,
13.9,
and
10.7
µ
g/
L
(
ppb),
respectively.
The
frequencies
of
detection
of
these
three
compounds
in
prepared
beverages
were
46%
(
28/
61),
43%
(
26/
61),
and
11%
(
7/
61),
respectively,
with
detection
limits
for
all
three
compounds
of
less
than
1
µ
g/
L
(
ppb).
In
contrast,
the
maximum
concentration
of
any
of
the
halogenated
organic
compounds
identified
in
mineral
water,
including
chloroform,
was
5.79
µ
g/
L
(
ppb).

McNeal
et
al.
(
1995)
examined
27
different
prepared
beverages
and
mineral
waters
in
the
United
States
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
at
detection
limits
of
0.1,
0.1,
and
0.2
ng/
g
(
ppb),
respectively.
Bromoform
was
not
detected
in
any
of
the
samples.
Bromodichloromethane
and
dibromochloromethane
were
detected
at
12
and
1
ng/
g
(
ppb),
respectively,
in
only
one
of
seven
types
of
mineral
and
sparkling
waters
examined.
The
positive
sample
was
the
only
sparkling
and
flavored
water
of
the
group,.
Bromodichloromethane
was
found
in
1
of
5
flavored
noncarbonated
beverages
examined,
a
fruit
drink,
at
a
concentration
of
5
ng/
g
(
ppb);
dibromochloromethane
was
not
detected
in
any
of
these
five
beverages.
Draft
­
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Not
Cite
or
Quote
February
20,
2003
IV
­
20
Bromodichloromethane
was
found
in
all
13
of
the
types
of
carbonated
soft
drinks
examined,
at
concentrations
ranging
from
1
to
4
ng/
g
(
ppb)
for
12
of
the
drinks
examined
and
at
12
ng/
g
(
ppb)
for
the
thirteenth.
Dibromochloromethane
was
detected
in
only
4
of
the
13
carbonated
soft
drinks
examined
at
levels
of
0.5
to
2
ng/
g
(
ppb).
None
of
the
brominated
trihalomethanes
was
detected
in
either
of
the
two
types
of
beer
examined.

McNeal
et
al.
(
1995)
also
examined
several
types
of
prepared
non­
beverage
foods
and
water
from
canned
vegetables
in
the
United
States
for
bromodichloromethane,
dibromochloromethane,
and
bromoform.
None
of
these
compounds
was
detected
in
any
of
the
samples.
The
foods
examined
included
two
types
of
canned
tomato
sauce,
canned
pizza
sauce,
canned
vegetable
juice,
vegetable
waters
from
two
types
of
canned
green
beans
and
one
type
of
sweet
corn,
duck
sauces,
beef
extract,
and
Lite
syrup
product.

The
U.
S.
Food
and
Drug
Administration
(
U.
S.
FDA,
2000)
has
analyzed
for
18
volatile
organic
hydrocarbons
(
VOCs),
including
bromodichloromethane
and
bromoform,
in
the
Total
Diet
Study
since
1995.
Bromodichloromethane
and
bromoform
were
analyzed
in
a
subset
of
70
food
items
in
14
Market
Baskets.
During
the
period
1995
to
1999,
bromodichloromethane
was
detected
in
one
sample
each
of
11
non­
beverage
food
items
(
sliced
bologna,
fried
eggs,
canned
pork
and
beans,
smooth
peanut
butter,
homemade
cornbread,
raw
orange,
canned
pineapple,
boiled
collards,
red
tomato,
green
pepper,
and
fast­
food
hamburger)
(
U.
S.
FDA,
2000).
The
detected
concentrations
ranged
from
10
to
16
ppb,
with
the
exception
of
fast
food
hamburger
which
contained
37
ppb.
Bromodichloromethane
was
detected
in
one
sample
of
bottled
apple
juice
at
a
concentration
of
33
ppb.
The
mean
detected
concentration
of
bromodichloromethane
in
three
samples
of
tap
water
was
18
ppb.
Dibromochloromethane
was
not
included
in
the
list
of
VOC
analytes
for
the
Total
Diet
Study.
Bromoform
was
listed
as
an
analyte,
but
no
detections
were
reported
in
the
data
summary
for
1991
to
1999.
The
detection
limits
for
bromodichloromethane
and
bromoform
were
not
reported.

Imaeda
et
al.
(
1994)
examined
bean
curd
commercially
available
in
Japan
for
trihalomethanes.
Neither
bromoform
nor
dibromochloromethane
were
detected
in
any
of
the
samples
at
a
detection
limit
of
0.1
ppb.
Bromodichloromethane
was
detected
in
6
of
10
samples
of
bean
curd
at
concentrations
ranging
from
1.2
to
5.2
ppb
and
in
1
of
10
samples
of
the
water
in
the
bean
curd
packages
at
5.2
ppb.

Kroneld
and
Reunanen
(
1990)
analyzed
for
brominated
trihalomethanes
in
samples
of
pasteurized
and
unpasteurized
cow's
milk
collected
in
Turku,
Finland.
The
average
concentration
of
bromodichloromethane
measured
in
pasteurized
milk
was
0.008
µ
g/
L
(
ppb)
(
range,
undetectable
to
0.03
µ
g/
L
(
ppb),
detection
limit
not
specified).
Dibromochloromethane
was
detected
in
only
one
sample
of
pasteurized
milk
at
5
µ
g/
L
(
ppb).
Traces
of
bromoform
were
detected
but
not
quantified.
Brominated
trihalomethanes
were
not
detected
in
unpasteurized
milk.
Their
presence
in
pasteurized
milk
was
considered
to
result
from
use
of
chlorinated
water
during
processing.

b.
Estimated
Dietary
Intake
Draft
­
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Cite
or
Quote
February
20,
2003
IV
­
21
Estimates
for
dietary
intake
of
brominated
trihalomethanes
by
residents
of
the
United
States
were
not
identified
in
the
materials
reviewed
for
this
document.
Furthermore,
information
on
the
levels
in
U.
S.
foods
is
too
limited
to
independently
calculate
a
reliable
estimate.
However,
the
available
data
suggest
that
the
concentrations
of
brominated
trihalomethanes
in
non­
beverage
foods
are
likely
low.
The
apparently
low
concentrations
of
brominated
trihalomethanes
in
nonbeverage
foods
are
consistent
with
the
physical
and
chemical
properties
of
these
compounds.
The
levels
of
individual
brominated
trihalomethanes
in
beverages
prepared
in
the
United
States
appear
to
be
less
than
or
about
equal
to
levels
measured
in
disinfected
surface
water.

Toyoda
et
al.
(
1990)
analyzed
the
dietary
intake
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
for
30
Japanese
housewives
in
Nagoya
and
Yokohama,
Japan.
Duplicate
portions
of
daily
meals
were
collected
for
three
consecutive
days
and
sampled
for
all
three
brominated
trihalomethanes.
The
types
of
food
consumed
were
not
reported.
This
omission
prevents
a
meaningful
comparison
of
the
studied
diet
to
that
consumed
by
the
U.
S.
population.
The
detection
limits
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
reported
to
be
0.1,
0.2,
and
0.5
ppb,
respectively.
The
concentration
of
bromodichloromethane
ranged
from
undetectable
to
1.7
ppb
(
average,
0.3
±
0.3
ppb
SD).
The
mean
daily
intake
of
bromodichloromethane
was
estimated
to
be
0.6
±
0.5
µ
g/
day.
The
concentration
of
dibromochloromethane
ranged
from
undetectable
to
0.6
ppb
(
average,
0.1
±
0.2
ppb),
and
the
mean
dietary
intake
was
estimated
to
be
0.3
±
0.3
µ
g/
day.
The
concentration
of
bromoform
ranged
from
undetectable
to
8.1
ppb
(
average,
0.5
±
1.3
ppb).
The
mean
dietary
intake
of
bromoform
was
estimated
to
be
0.9
±
1.3
µ
g/
day.

Brominated
trihalomethanes
have
been
detected
in
a
number
of
beverages.
In
conducting
an
exposure
assessment,
the
potential
exposures
from
drinking
prepared
beverages
would
not
be
added
to
the
default
assumption
of
an
adult
consuming
2
liters
of
drinking
water
per
day.
Instead,
the
prepared
beverages
would
be
considered
part
of
the
2
liters
of
fluid
intake
per
person
per
day.

2.
Air
Intake
a.
Concentrations
in
Outdoor
Air
Brominated
trihalomethanes
are
usually
found
in
outdoor
air
at
low
concentrations
when
all
data,
including
nondetects,
are
considered.
Brodzinsky
and
Singh
(
1983)
reviewed,
summarized,
and
critically
evaluated
existing
data
for
brominated
trihalomethane
concentrations
in
ambient
outdoor
air
for
several
urban/
suburban
or
source
dominated
locations
across
the
United
States
(
Table
IV­
8).
No
concentration
data
were
available
for
rural
or
remote
areas.
The
authors
reported
mean,
median,
first
and
third
quartile
values,
and
minimum
and
maximum
values
by
city.
In
addition,
they
reported
the
same
measures
when
the
data
were
grouped
by
type
of
location
(
i.
e.,
urban/
suburban
or
source
dominated),
and
when
all
data
were
combined.
Ambient
air
concentrations
were
reported
for
bromodichloromethane
at
Magnolia,
AR,
El
Dorado,
TX,
Chapel
Hill,
NC,
and
Beaumont,
TX.
Bromodichloromethane
was
detected
at
mean
concentrations
of
0.76
ppt,
1.40
ppt,
120
ppt,
and
180
ppt
for
those
four
cities,
respectively,
Draft
­
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Cite
or
Quote
February
20,
2003
IV
­
22
where
ppt
is
expressed
as
parts
per
trillion
by
volume.
Dibromochloromethane
was
detected
in
the
air
samples
from
Magnolia,
AR,
El
Dorado,
TX,
Chapel
Hill,
NC,
Beaumont
TX,
and
Lake
Charles,
LA
at
mean
concentrations
of
0
ppt,
0.48
ppt,
14
ppt,
14
ppt,
and
19
ppt,
respectively.
Bromoform
was
detected
in
air
samples
from
Magnolia,
AR,
El
Dorado,
TX,
and
Lake
Charles,
LA,
at
concentrations
of
1.5
ppt,
0.81
ppt,
and
50
ppt,
respectively.
Air
concentration
data
from
these
sites
were
combined
for
additional
statistical
analysis.
The
study
authors
indicated
that
a
value
of
0.0
was
entered
for
samples
below
the
detection
limit.
Mean
(
±
standard
deviation)
outdoor
air
concentrations
in
urban/
suburban
and
source
dominated
locations,
respectively,
were
160
±
29
ppt
and
1.2
±
0.4
ppt
for
bromodichloromethane;
15
±
4
ppt
and
0.28
±
0.67
ppt
for
dibromochloromethane;
and
50
±
29
ppt
and
1.1
±
2.1
ppt
for
bromoform.
Brodzinsky
and
Singh
(
1983)
also
calculated
overall
(
grand)
means
based
on
data
from
all
sites.
Grand
mean
values
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
110
ppt
(
n
=
26,
with
one
nondetect),
3.8
ppt
(
n
=
89,
with
63
nondetects),
and
3.6
ppt
(
n
=
78,
with
60
nondetects),
respectively.
When
expressed
on
a
µ
g/
m3
basis,
the
corresponding
mean
values
for
bromodichloromethane,
dibromochloromethane,
and
bromoform
are
0.74
µ
g/
m3,
0.032
µ
g/
m3,
and
0.037
µ
g/
m3.
Draft
­
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Not
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or
Quote
February
20,
2003
IV
­
23
Table
IV­
8.
Selected
Concentration
Data
for
Individual
Brominated
Trihalomethanes
(
ppt)
in
Outdoor
Air
as
Summarized
in
Brodzinsky
and
Singh
(
1983)
a,
b
City
n
Nondetects
Mean
(
Std
dev.)
Median
3rd
Quartile
Maximum
Reference
Bromodichloromethane
Individual
Sites
Beaumont,
TX
11
0
180
(
100)
180
180
180
Wallace
(
1981)

Chapel
Hill,
NC
6
0
120
(
210)
120
120
120
Wallace
(
1981)

El
Dorado,
AR
7
1
1.4
(
0.35)
1.6
1.6
1.6
Pellizzari
and
Bunch
(
1979)

Magnolia,
AR
2
0
0.76
(
0.0)
0.0
0.0
0.76
Pellizzari
and
Bunch
(
1979)

Totals
Urban/
Suburban
17
0
160
(
29)
180
180
180
­

Source
Areas
9
1
1.2
(
0.41)
1.6
1.6
1.6
­

Grand
totals
26
1
110
(
82)
120
180
180
­

Dibromochloromethane
Individual
Sites
Beaumont,
TX
11
0
14
(
0.0)
14
14
14
Wallace
(
1981)

Chapel
Hill,
NC
6
0
14
(
0.0)
14
14
14
Wallace
(
1981)

El
Dorado,
AR
40
35
0.48
(
0.82)
0.0
0.82
2.5
Pellizzari
et
al.
(
1978)

Lake
Charles,
LA
4
0
19
(
9.6)
21
27
27
Pellizzari
(
1979)

Magnolia,
AR
28
28
0.0
(
0.0)
0.0
0.0
0.0
Pellizzari
et
al.
(
1978)
Table
IV­
8
(
cont.)

City
n
Nondetects
Mean
(
Std
dev.)
Median
3rd
Quartile
Maximum
Reference
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
24
Dibromochloromethane
(
cont.)

Totals
Urban/
Suburban
21
0
15
(
4.2)
14
14
27
­

Source
Areas
68
63
0.28
(
0.67)
0.0
0.0
2.5
­

Grand
Totals
89
63
3.8
(
6.7)
0.0
2.5
27
­

Bromoform
Individual
Sites
El
Dorado,
AR
46
35
0.81
(
0.95)
0.43
1.3
2.7
Pellizzari
et
al.
(
1978)

Pellizzari
and
Bunch
(
1979)

Lake
Charles,
LA
4
0
50
(
29)
62
68
71
Pellizzari
(
1979)

Magnolia,
AR
28
25
1.5
(
3.2)
0.0
0.29
8.3
Pellizzari
et
al.
(
1978)

Totals
Urban/
Suburban
4
0
50
(
29)
62
68
71
­

Source
Areas
74
60
1.1
(
2.1)
0.0
1.3
8.3
­

Grand
Totals
78
60
3.6
(
12)
0.0
1.5
71
­

a
Includes
only
data
considered
to
be
of
adequate,
good,
or
excellent
quality
by
the
study
authors.

b
Concentrations
are
reported
as
parts
per
trillion
by
volume
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
25
Shikiya
et
al.
(
1984)
analyzed
ambient
air
samples
collected
at
four
urban/
industrial
locations
in
the
California
South
Coast
Air
Basin
from
November
1982
to
December
1983
for
the
presence
of
halogenated
hydrocarbons.
Data
for
bromodichloromethane,
dibromochloromethane
and
bromoform
were
included
in
this
analysis.
The
sampling
locations
were
El
Monte,
downtown
Los
Angeles,
Dominguez,
and
Riverside.
The
air
samples
were
analyzed
using
gas
chromatography
with
detection
by
electron
capture.
The
quantitation
limit,
defined
as
a
level
10
times
greater
than
the
noise
level,
was
10
ppt
by
volume
for
all
three
brominated
trihalomethanes.
The
detection
limit
was
defined
as
three
times
the
noise
level.
Summary
data
for
each
compound
included
monthly
means
and
composite
means.
The
monthly
means
were
calculated
as
the
average
of
all
data
at
a
site
that
were
above
the
quantitation
limit
for
a
single
month;
samples
with
concentrations
below
the
limit
of
detection
were
not
included
in
the
calculations.
The
composite
means
were
calculated
as
the
average
value
of
all
data
for
each
compound
above
the
quantitation
limit
at
each
site.
Most
data
in
this
report
were
presented
graphically.
A
few
additional
details
were
presented
in
a
short
summary
statement
for
each
chemical.
Thirty­
five
percent
of
the
samples
had
bromodichloromethane
levels
above
the
quantitation
limit
of
10
ppt
(
0.067
µ
g/
m3).
Peaks
in
the
concentration
of
bromodichloromethane
were
observed
at
various
sites
in
June
and
July,
with
downtown
Los
Angeles
and
Dominguez
registering
the
highest
monthly
means
of
approximately
30
ppt
(
0.20
µ
g/
m3).
The
highest
reported
concentration
was
40
ppt
(
0.27
µ
g/
m3).
The
highest
composite
mean
of
100
ppt
(
0.67
µ
g/
m3)
for
bromodichloromethane
was
observed
at
El
Monte.
In
comparison,
the
remaining
three
locations
had
a
composite
mean
of
20
ppt
(
0.08
µ
g/
m3).
For
dibromochloromethane,
only
seventeen
percent
of
the
samples
had
levels
above
the
quantitation
limit
of
10
ppt
(
0.085
µ
g/
m3).
The
highest
reported
concentration,
monthly
mean,
and
mean
composite
for
dibromochloromethane
were
290
ppt
(
2.5
µ
g/
m3),
280
ppt
(
2.4
µ
g/
m3),
and
50
ppt
(
0.43
µ
g/
m3),
respectively;
all
were
recorded
in
downtown
Los
Angeles
in
June.
Only
two
monthly
means
were
above
160
ppt;
the
remainder
of
the
monthly
means
were
below
60
ppt.
For
bromoform,
thirty­
one
percent
of
the
samples
had
concentrations
above
the
quantitation
limit
of
10
ppt
(
0.10
µ
g/
m3).
Peaks
in
the
concentration
of
bromoform
were
observed
at
various
sites
in
May
and
June,
with
the
downtown
Los
Angeles
site
registering
the
highest
composite
mean
(
40
ppt;
0.41
µ
g/
m3)
and
the
highest
monthly
mean
(
310
ppt;
3.2
µ
g/
m3)
in
June
1983.
Only
two
monthly
means
were
greater
than
160
ppt;
the
remainder
of
the
monthly
means
were
below
60
ppt.

Atlas
and
Schauffler
(
1991)
collected
replicate
air
samples
at
various
locations
on
the
Island
of
Hawaii
during
a
month­
long
field
experiment
to
test
an
analytical
method
for
determining
halocarbons
in
ambient
air.
Dibromochloromethane
was
found
at
a
mean
level
of
0.27
ppt,
and
bromoform
was
found
at
a
mean
concentration
of
1.9
ppt.
Information
on
sample
size
and
detection
limit
were
not
provided
in
the
secondary
source
that
reported
this
study
(
U.
S.
EPA
1994b).

Wallace
et
al.
(
1982)
conducted
a
pilot
study
designed
to
field
test
personal
air­
quality
monitoring
methods.
Personal
air
samples
were
collected
from
students
at
two
universities:
Lamar
University,
Texas,
located
near
a
petrochemical
manufacturing
area,
and
the
University
of
North
Carolina
(
UNC),
located
in
a
nonindustrialized
area.
The
samples
were
analyzed
for
a
number
of
volatile
organic
compounds,
including
brominated
trihalomethanes.
Draft
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IV
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26
Bromodichloromethane
was
detected
in
64%
of
personal
air
samples
from
11
Lamar
students,
with
a
mean
of
1.23
µ
g/
m3
(
0.18
ppb),
a
median
of
1
µ
g/
m3
(
0.15
ppb),
and
a
range
of
0.12
 
3.72
µ
g/
m3
(
0.018
 
0.56
ppb).
The
limit
of
detection
was
0.24
µ
g/
m3
(
0.036
ppb).
At
UNC,
17%
of
the
samples
from
6
students
had
detectable
levels
of
bromodichloromethane.
Concentrations
ranged
from
0.12
 
4.36
µ
g/
m3
(
0.017
 
0.65
ppb)
(
mean,
0.83
µ
g/
m3
(
0.12
ppb);
median,
0.12
µ
g/
m3
(
0.017
ppb)).
Based
on
the
above
information,
the
average
daily
intake
of
bromodichloromethane
from
air
using
an
inhalation
rate
of
20
m3/
day
was
estimated
to
be
25
µ
g/
day
for
Lamar
students
and
17
µ
g/
day
for
UNC
students.
Dibromochloromethane
was
not
present
above
0.12
µ
g/
m3
(
0.018
ppb)
at
either
site.

b.
Concentrations
in
Indoor
Air
Relatively
few
studies
have
reported
the
concentrations
of
trihalomethanes
in
the
indoor
air
of
homes.
Kostiainen
(
1995)
identified
over
200
volatile
organic
compounds
in
indoor
air
of
26
houses
identified
by
residents
as
causing
symptoms
such
as
headache,
nausea,
irritation
of
the
eyes,
drowsiness,
and
fatigue.
Bromoform
was
detected
at
low
(
unspecified)
levels
in
54
percent
of
the
homes,
and
no
mention
was
made
of
dibromochloromethane
or
bromodichloromethane.

Weisel
et
al.
(
1999)
measured
brominated
trihalomethane
concentrations
in
indoor
air
in
New
Jersey
residences
selected
to
examine
low
and
high
levels
of
drinking
water
contamination
with
trihalomethanes.
Descriptive
statistics
for
trihalomethane
concentration
in
water
were
provided
for
the
combined
high
and
low
concentration
groups,
but
not
for
the
individual
categories.
One
valid
15­
minute
air
sample
was
collected
at
each
of
48
residences.
The
indoor
air
concentrations
of
bromodichloromethane
averaged
0.38
±
0.82
(
SD)
µ
g/
m3
(
0.057
±
0.12
ppb)
and
0.75
±
0.96
µ
g/
m3
(
0.11
±
0.14
ppb)
from
the
low
and
high
water
concentration
groups,
respectively.
The
detection
frequencies
were
12/
25
and
16
/
23
in
the
low
and
high
water
concentration
groups,
respectively.
The
indoor
air
concentrations
of
dibromochloromethane
averaged
0.44
±
0.95
µ
g/
m3
(
0.052
±
0.11
ppb)
and
0.53
±
0.84
µ
g/
m3
(
0.062
±
0.09
ppb)
from
the
low
and
high
water
concentration
groups
with
detection
frequencies
of
5/
25
and
7/
23,
respectively.
For
bromoform,
the
average
concentrations
from
the
low
and
high
water
concentration
groups
were
0.29
±
0.93
µ
g/
m3
(
0.028
±
0.089
ppb)
and
0.35
±
0.94
µ
g/
m3
(
0.034
±
0.091
ppb),
with
detection
frequencies
of
8/
25
and
4/
23,
respectively.
It
was
not
clear
whether
the
averages
were
based
on
all
measured
samples
or
only
those
samples
that
were
above
the
detection
limit
for
each
compound.

Kerger
et
al
(
2000)
evaluated
the
transfer
of
bromodichloromethane
and
dibromochloromethane
to
indoor
air
in
bathrooms
during
showering
and
bathing
in
homes
supplied
with
chlorinated
tap
water.
The
test
sites
were
three
urban
homes
containing
three
bedrooms,
a
full
bath,
and
approximately
1000
square
feet
of
living
space.
The
compounds
were
simultaneously
measured
in
hot
and
cold
tap
water
(
drawn
from
the
kitchen
sink)
and
in
the
shower/
bath
enclosure
and
bathroom
vanity
area.
Three
shower
protocols
were
examined:
6.8
min
unventilated
shower;
12
min
unventilated
shower
and
6.8
min
ventilated
shower.
Water
flow
rate
and
temperature
were
monitored
but
not
controlled.
Airborne
vapor
samples
were
captured
by
Summa
canister
and
measured
by
gas
chromatography
using
electron
capture
detection
Draft
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IV
­
27
according
to
U.
S.
EPA
method
TO­
14.
Air
samples
were
collected
before,
during
and
after
the
water
use
event,
for
a
total
of
16
showers
and
7
baths.
Data
for
several
events
were
eliminated
because
of
technical
difficulties.
For
all
shower
protocols
combined
(
n
=
12),
the
increase
in
average
airborne
concentration
(
±
standard
error),
expressed
as
µ
g/
m3,
in
shower
enclosure
or
bathroom
air
per
µ
g/
L
in
water,
was
1.8
±
0.3
for
bromodichloromethane
and
0.5
±
0.1
for
dibromochloromethane.
For
baths
(
n
=
4),
the
average
concentration
increase
during
the
bath
was
0.59
±
0.21
for
bromodichloromethane
and
0.15
±
0.05
for
dibromochloromethane.
The
relative
contribution
of
each
chemical
was
consistent
with
the
relative
concentration
in
water
and
its
chemical
and
physical
properties.
The
average
exposures
measured
in
this
study
were
approximately
30%
lower
than
results
reported
by
other
investigators
using
EPA
analytical
methods
when
data
were
normalized
for
water
concentration,
flow
rate,
shower
volume,
and
duration.
This
difference
may
have
resulted
from
differences
in
the
air
exchange
rate
between
residential
showers
and
laboratory
test
showers.
These
data
are
not
adequate
for
characterizing
levels
of
individual
brominated
trihalomethanes
in
the
home
because
the
measurements
targeted
a
specific
area
of
the
residences
and
the
sample
size
consisted
of
only
three
homes.

c.
Estimates
of
Exposure
from
Air
The
data
available
for
occurrence
of
brominated
trihalomethanes
in
air
do
not
permit
calculation
of
a
nationally
aggregated
intake
estimate
for
the
U.
S.
general
population.
To
accurately
estimate
total
daily
inhalation
exposures,
factors
including
location
and
season,
the
fraction
of
time
spent
indoors
compared
with
outdoors,
potential
exposures
of
individuals
while
showering
or
bathing,
potential
exposure
from
volatilization
of
brominated
trihalomethanes
during
other
household
activities
(
e.
g.,
use
of
dishwashers,
toilet
flushing),
exposures
of
individuals
who
spend
large
amounts
of
time
at
indoor
pools,
and
potential
for
occupational
exposures
(
e.
g.,
for
laundromat
or
sewage
treatment
plant
workers)
require
consideration.
Although
the
existing
data
do
not
permit
such
a
refined
analysis,
they
may
be
used
to
roughly
estimate
intake
from
air.
Based
on
the
grand
means
calculated
for
multiple
sampling
locations
by
Brodzinsky
and
Singh
(
1983),
exposure
to
bromodichloromethane,
dibromochloromethane
and
bromoform
resulting
from
inhalation
of
outdoor
air
can
be
roughly
estimated
assuming
an
inhalation
rate
of
20
m3/
day,
100%
absorption,
and
exposure
to
outdoor
air
for
a
full
24
hours
per
day.
Using
the
mean
ambient
air
concentration
of
110
ppt
(
0.74
µ
g/
m3)
by
volume
for
all
sites
reported
in
Brodzinsky
and
Singh
(
1983),
the
daily
intake
of
bromodichloromethane
from
outdoor
air
would
be
21
µ
g/
day.
Assuming
a
mean
air
concentration
of
3.8
ppt
(
0.032
µ
g/
m3)
for
dibromochloromethane,
daily
intake
would
be
0.64
µ
g/
day,.
Assuming
a
mean
air
concentration
of
3.6
ppt
(
0.037
µ
g/
m3)
by
volume
or
bromoform,
the
daily
intake
would
be
0.74
µ
g/
day.
Because
these
estimates
are
based
on
data
from
urban/
suburban
and
industrial
sites
only,
they
may
represent
high
end
exposures.

Adequate,
nationally
aggregated
occurrence
data
are
not
available
for
calculating
intake
of
brominated
trihalomethanes
from
indoor
air.
The
indoor
air
concentrations
measured
by
Weisel
et
al.
(
1999)
were
not
used
for
intake
calculations
because
it
could
not
be
determined
how
the
means
for
each
compound
were
calculated
(
i.
e.,
whether
all
measurements
were
averaged
or
only
those
Draft
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February
20,
2003
IV
­
28
above
the
detection
limit).
In
addition,
the
data
were
based
on
a
single
15
minute
air
sample
collected
from
each
of
48
homes
located
in
a
single
state.

While
brominated
trihalomethane
concentrations
might
be
expected
to
be
higher
in
indoor
air
than
in
outdoor
air
due
to
confined
space
and
additional
indoor
air
sources
(
e.
g.
volatilization
from
showering,
baths,
and
other
household
activities),
the
available
data
do
not
allow
such
a
comparison.

Based
on
data
from
personal
air
monitors,
Wallace
et
al.
(
1982)
estimated
daily
inhalation
of
bromodichloromethane
to
be
25
µ
g/
day
for
11
students
attending
a
university
located
near
a
petrochemical
manufacturing
area
and
17
µ
g/
day
for
6
students
attending
a
university
in
a
nonindustrialized
area..
The
personal
air
monitors
registered
bromodichloromethane
from
both
indoor
(
with
the
exception
of
showering
and
bathing)
and
outdoor
exposures.
Dibromochloromethane
was
not
detected
and
no
data
were
available
for
bromoform.

3.
Concentrations
and
Exposures
Associated
with
Swimming
Pools
and
Hot
Tubs
Numerous
studies
have
reported
data
for
concentrations
of
brominated
trihalomethanes
and
exposures
associated
with
swimming
pools
and
hot
tubs.
Exposure
of
swimmers
or
hot
tub
users
to
brominated
trihalomethanes
may
result
from
dermal,
ingestion,
and
inhalation
exposure.
When
evaluating
these
data,
it
is
important
to
note
that
additional
disinfectants
are
routinely
added
to
water
contained
in
swimming
pools
and
hot
tubs;
therefore,
the
levels
of
brominated
trihalomethanes
present
may
not
be
representative
of
those
in
tap
water.

Armstrong
and
Golden
(
1986)
measured
bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
in
the
water
and
surrounding
air
of
four
indoor
swimming
pools,
five
outdoor
swimming
pools,
and
four
hot
tubs.
Concentrations
in
air
were
measured
two
centimeters
from
the
water
surface.
The
bromodichloromethane
concentrations
of
water
in
the
outdoor
pools
ranged
from
1
to
72
µ
g/
L
(
ppb)
(
mean,
33
µ
g/
L).
Levels
in
the
indoor
pools
ranged
from
1
to
90
µ
g/
L
(
ppb)
(
mean,
16
µ
g/
L).
The
levels
of
bromodichloromethane
in
the
hot
tubs
ranged
from

0.1
to
105
µ
g/
L
(
ppb)
(
mean,
17
µ
g/
L).
Means
and
ranges
of
the
bromodichloromethane
concentration
two
meters
above
the
water
surface
for
outdoor
pools,
indoor
pools,
and
hot
tubs,
respectively,
were:
<
0.1
µ
g/
m3
(<
0.015
ppb)
(
range
not
reported),
1.7
µ
g/
m3
(
0.25
ppb)
(
range
<
0.1
 
10
µ
g/
m3
(
0.015
 
1.5
ppb)),
and
1.4
µ
g/
m3
(
0.21
ppb)
(
range
<
0.1
 
10
µ
g/
m3
(
0.015
 
1.5
ppb)).
The
dibromochloromethane
concentration
of
water
in
the
outdoor
pools
ranged
from
<
0.1
to
8
µ
g/
L
(
ppb)
(
mean,
4.2

g/
L
(
ppb)).
Levels
in
the
indoor
pools
ranged
from
0.3
to
30
µ
g/
L
(
ppb)
(
mean,
9.5
µ
g/
L
(
ppb)).
The
level
of
dibromochloromethane
in
the
hot
tubs
ranged
from

0.1
to
48
µ
g/
L
(
ppb)
(
mean,
14.4
µ
g/
L
(
ppb)).
Means
and
ranges
of
the
dibromochloromethane
concentration
two
meters
above
the
water
surface
for
outdoor
pools,
indoor
pools,
and
hot
tubs,
respectively,
were:
<
0.1
µ
g/
m3
(<
0.01
ppb)
(
range
not
reported),
0.9
µ
g/
m3
(
0.11
ppb)
(<
0.1
 
5
µ
g/
m3
(
0.012
 
0.59
ppb)),
and
0.7
µ
g/
m3
(
0.08
ppb)
(<
0.1
 
5
µ
g/
m3
(
0.012
 
0.59
ppb)).
The
mean
bromoform
concentration
in
the
outdoor
pools
was
less
than
0.1
µ
g/
L
(
ppb).
Levels
in
the
indoor
pools
ranged
from
less
than
0.1
to
20
µ
g/
L
(
ppb)
(
mean,
6
µ
g/
L
(
ppb)).
The
levels
of
bromoform
in
the
hot
tubs
ranged
from
less
than
0.1
to
62
µ
g/
L
Draft
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February
20,
2003
IV
­
29
(
ppb)
(
mean,
13
µ
g/
L
(
ppb)).
Means
and
ranges
of
the
bromoform
concentration
two
meters
above
the
water
surface
for
outdoor
pools,
indoor
pools,
and
hot
tubs,
respectively,
were:
<
0.1
µ
g/
m3
(<
9.7
ppt)
(
range
not
reported),
9
µ
g/
m3
(
870
ppt)
(<
0.1
 
14
µ
g/
m3
(
9.7
 
1360
ppt)),
and
8
µ
g/
m3
(
770
ppt)
(<
0.1
 
14
µ
g/
m3
(
9.7
 
1360
ppt)).

Cammann
and
Hübner
(
1995)
compared
concentrations
of
trihalomethanes
in
swimmers'
and
bath
attendants'
blood
and
urine
before
and
after
swimming
or
working
in
indoor
swimming
pools.
Water
and
air
concentrations
were
measure
in
different
locations
in
the
pool
environment.
The
purpose
was
to
determine
whether
blood
levels
of
trihalomethanes
would
reflect
inhalation
exposure
to
trihalomethanes
in
the
pool
environment
and
whether
those
compounds
also
would
appear
in
urine.
Measured
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
samples
of
swimming
pool
waters
collected
at
a
depth
of
10
to
20
cm
were
0.69
to
5.64
µ
g/
L
(
ppb),
0.03
to
6.51
µ
g/
L
(
ppb),
and
0.14
to
2.32
µ
g/
L
(
ppb)
for
the
three
compounds,
respectively.
Averages
(
±
SD)
of
the
10
pool
water
measurements
presented
in
Table
1
of
the
report
were
2.12
±
1.52
µ
g/
L
(
ppb),
1.11
±
2.07
µ
g/
L
(
ppb),
and
0.42
±
0.73
µ
g/
L
(
ppb)
for
bromodichloromethane,
dibromochloromethane,
and
bromoform,
respectively.
Average
(
±
1
SD)
concentrations
in
the
four
air
samples
taken
(
location
of
sampling
not
specified)
were
15.4
±
7.36
µ
g/
m3
(
2.30
±
1.10
ppb),
1.94
±
1.01
µ
g/
m3
(
0.228
±
0.119
ppb),
and
below
the
quantitation
limit
(
QL)
(
not
specified,
although
probably
0.02
ppb)
for
bromodichloromethane,
dibromochloromethane,
and
bromoform,
respectively.

Measurements
of
bromodichloromethane
in
8
bath
attendants'
blood
before
their
shifts
ranged
from
below
QL
for
12/
18
measurements
(
67%)
to
0.1
µ
g/
L
(
ppb)
(
Camman
and
Hübner,
1995).
After
their
shifts,
the
concentrations
ranged
from
below
QL
in
7/
18
measurements
(
39%)
to
0.6

g/
L
(
ppb).
Similarly,
measurements
of
bromodichloromethane
in
swimmers'
blood
was
higher
after
than
before
swimming.
Before
swimming,
blood
concentrations
of
bromodichloromethane
ranged
from
less
than
the
QL
in
10/
20
(
50%)
swimmers
to
0.2
µ
g/
L
(
ppb);
while
after
swimming,
blood
concentrations
were
above
the
QL
in
all
20
swimmers,
ranging
from

0.02
to
0.4

g/
L
(
ppb)
in
19
of
the
swimmers.
The
twentieth
swimmer
had
a
blood
concentration
of

1.5

g/
L
(
ppb).
For
all
but
two
of
the
swimmers,
blood
concentrations
of
bromodichloromethane
had
dropped
below
the
QL
by
the
next
day
(
values
for
the
other
two
swimmers
were
less
than
0.1

g/
L
(
ppb)).
Dibromochloromethane
and
bromoform
were
not
detected
in
the
blood
of
either
the
bath
attendants
or
swimmers.
None
of
the
brominated
trihalomethanes
were
detected
in
the
urine
of
the
study
subjects.
Thus,
only
exposure
to
bromodichloromethane
by
inhalation
(
bath
attendants)
or
inhalation,
dermal
absorption,
and
ingestion
(
swimmers)
is
reflected
in
increased
blood
levels
of
the
compound.
Blood
levels
of
bromodichloromethane
usually
returned
to
pre­
exposure
levels
within
24
hours
after
the
exposure
.

Aggazzotti
et
al.
(
1998)
evaluated
concentrations
of
trihalomethanes
in
the
blood
and
breath
of
five
competitive
swimmers
regularly
training
in
an
indoor
swimming
pool
in
Italy.
The
group
included
three
males
and
two
females
between
the
ages
of
17
and
21
years.
All
were
nonsmokers
Concurrent
sampling
of
blood,
alveolar
air,
and
environmental
air
occurred
at
five
times
for
each
of
four
sessions:
(
a)
at
the
University
Department
two
hours
before
arriving
at
the
pool,
Draft
­
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February
20,
2003
IV
­
30
(
b)
after
one
hour
sitting
near
the
edge
of
the
pool,
(
c),
after
one
hour
of
swimming,
(
d)
back
at
the
University
one
hour
after
swimming
ended,
and
(
e)
at
the
University
1.5
hr
after
swimming
ended.
While
bromodichloromethane
and
dibromochloromethane
were
always
found
in
water
and
environmental
air
samples
at
the
pool
immediately
before
and
after
the
1­
hr
swimming
session,
bromoform
was
rarely
detected
in
the
indoor
pool
air.
None
of
the
three
brominated
trihalomethanes
were
detected
in
the
air
at
the
University
Department
or
in
the
alveolar
air
of
the
swimmers
at
the
Department
two
hours
before
arriving
at
the
pool.
At
the
pool,
prior
to
the
swimming
session,
the
means
(
±
SD)
of
the
four
measured
ambient
air
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
were10.5
±
3.1
µ
g/
m3
(
1.6
±
0.46
ppb;
4
detects),
5.2
±
1.5
µ
g/
m3
(
0.61
±
0.17
ppb;
4
detects),
and
1
detect
of
0.2
µ
g/
m3
(
0.02
ppb),
respectively.
At
the
pool,
just
after
the
1­
hr
swimming
session,
the
means
(
±
SD)
of
the
four
measured
ambient
air
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
20.0
±
4.1
µ
g/
m3
(
2.99
±
4.1
ppb;
4
detects),
11.4
±
2.1
µ
g/
m3
(
1.34
±
0.23
ppb;
4
detects),
and
1
detect
of
0.2
µ
g/
m3
(
0.02
ppb),
respectively.

Concentrations
of
bromodichloromethane
and
dibromochloromethane
in
the
alveolar
air
of
the
swimmers
before
and
after
the
swimming
session
indicated
inhalation
uptake
of
both
compounds
(
Aggazzotti
et
al.,
1998).
At
the
pool,
prior
to
the
swimming
session,
the
means
(
±
SD)
of
the
20
measured
alveolar
air
concentrations
(
5
swimmers
assessed
at
each
of
4
sessions)
of
bromodichloromethane
and
dibromochloromethane
were
2.7
±
1.2
µ
g/
m3
(
0.40
±
0.18
ppb)
and
0.8
±
0.8
µ
g/
m3
(
0.09
±
0.09
ppb),
respectively.
Bromoform
was
not
detected
in
any
of
the
20
samples.
At
the
pool,
after
the
1­
hr
swimming
session,
the
means
(
±
SD)
of
the
alveolar
air
concentrations
of
bromodichloromethane
and
dibromochloromethane
were
6.5
±
1.3
µ
g/
m3
(
0.97
±
0.19
ppb)
and
1.4
±
0.9
µ
g/
m3
(
0.16
±
0.11
ppb),
respectively.
Bromoform
was
not
detected
in
any
of
the
20
samples.
Blood
levels
of
bromodichloromethane
and
dibromochloromethane
before
and
after
swimming,
on
the
other
hand,
were
below
detection
limits
in
most
samples,
and
hence
showed
no
trends.

Aggazzotti
et
al.
(
1998)
estimated
uptake
of
the
trihalomethanes
of
the
resting
and
active
swimmers
using
the
following
assumptions.
At
rest,
the
pulmonary
ventilation
rate
of
the
women
was
6
liters
per
minute
(
L/
min)
while
that
of
men
was
7.5
L/
min.
During
swimming,
the
ventilation
rate
of
the
women
was
25
L/
min
while
that
of
the
men
36
L/
min.
The
estimated
uptake
rates
of
bromodichloromethane
for
the
five
swimmers
at
rest
ranged
from
2.8
to
3.7

g
per
hour
(

g/
h),
with
a
mean
value
of
3.3
±
0.41(
SD)

g/
h
for
the
three
males
and
two
females
combined.
The
estimated
uptake
rates
for
the
same
individuals
actively
swimming
were
20
to
30

g/
h,
with
a
mean
value
of
26
±
5.1

g/
h.
The
estimated
uptake
rates
of
dibromochloromethane
for
the
five
swimmers
at
rest
ranged
from
1.5
to
2.0

g/
h,
with
a
mean
value
of
1.8
±
0.23

g/
h.
The
estimated
uptake
rates
during
swimming
increased
to
between
14
and
22

g/
h,
with
a
mean
value
of
18
±
3.6

g/
h.
Occurrence
of
dermal
uptake
was
acknowledged
but
not
estimated.

Lindstrom
et
al.
(
1997)
also
assessed
exposure
of
two
competitive
swimmers
to
bromodichloromethane
during
training
sessions
at
an
indoor
pool.
The
indoor
pool
air
concentrations
of
bromodichloromethane
collected
over
60­
and
119­
minute
intervals
were
2.76
and
3.02
µ
g/
m3
(
0.41
and
0.45
ppb),
respectively.
Breath
samples
were
collected
from
the
Draft
­
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February
20,
2003
IV
­
31
swimmers
before,
during,
and
for
3
hours
after
a
training
workout.
Breath
samples
collected
during
the
workout
demonstrated
a
rapid
uptake
of
bromodichloromethane
to
maximum
alveolar
concentrations
of
5
to
6
µ
g/
m3
(
0.7
to
0.9
ppb),
which
are
higher
than
the
ambient
air
concentrations.
The
authors
concluded
that
significant
(
80%
of
total
exposure)
dermal
absorption
of
the
related
trihalomethane
chloroform
from
water
was
occurring,
but
did
not
estimate
the
extent
of
dermal
uptake
for
bromodichloromethane.

4.
Soil
Concentrations
and
Exposure
Data
on
the
concentration
of
brominated
trihalomethanes
in
soil
were
not
available
in
the
materials
reviewed
for
this
document.
Based
on
the
measured
Henry's
Law
constant
and
vapor
pressure
of
the
individual
compounds,
volatilization
from
both
wet
and
dry
soil
surfaces
should
be
relatively
rapid
(
U.
S.
EPA
1987).
Therefore.
exposure
from
soil
ingestion
is
not
considered
to
be
a
significant
route
for
exposure
to
the
brominated
trihalomethanes.

C.
Overall
Exposure
The
RSC
(
relative
source
contribution)
is
the
percentage
of
total
daily
exposure
that
is
attributable
to
tap
water
when
all
potential
sources
are
considered
(
e.
g.,
air,
food,
soil,
and
water).
Ideally,
the
RSC
is
determined
quantitatively
using
nationwide,
central
tendency
and/
or
high­
end
estimates
of
exposure
from
each
relevant
medium.
In
the
absence
of
such
data,
a
default
RSC
ranging
from
20%
to
80%
may
be
used.

The
RSC
used
in
the
current
and
previous
drinking
water
regulations
for
dibromochloromethane
is
80%.
This
value
was
established
by
use
of
a
screening
level
approach
to
estimate
and
compare
exposure
to
dibromochloromethane
from
various
sources.
Information
considered
for
during
this
process
is
summarized
in
Appendix
C.
The
use
of
the
80%
value
for
the
RSC
for
dibromochloromethane
is
supported
by
limited
use
of
this
chemical
in
industrial
applications
with
potential
for
direct
release
to
the
environment.
The
use
of
the
80%
value
is
further
supported
by
apparently
low
concentrations
in
foods
and
soils
and
the
potential
for
human
exposure
to
dibromochloromethane
in
tap
water
via
three
exposure
routes:
1)
ingestion
as
drinking
water;
2)
inhalation
of
volatilized
dibromochloromethane
during
use
of
tap
water
for
household
activities;
and
3)
by
dermal
exposure
during
showering,
bathing,
or
other
activities.
The
available
data
for
concentrations
of
outdoor
air
and
food,
although
limited,
suggest
that
exposures
via
these
routes
are
likely
to
be
low
when
compared
to
water.

Parallel
RSC
calculations
were
not
performed
for
bromodichloromethane
and
bromoform.
The
EPA
has
set
the
regulatory
level
for
these
chemicals
in
drinking
water
at
zero
because
it
has
been
determined
that
they
are
probable
human
carcinogens.
Therefore,
determination
of
an
RSC
is
not
relevant
for
these
chemicals
because
it
is
the
Agency's
policy
to
perform
RSC
analysis
only
for
noncarcinogens.
Draft
­
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or
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February
20,
2003
IV
­
32
D.
Body
Burden
1.
Blood
and
Breath
Levels
Barkley
et
al.
(
1980)
analyzed
blood
samples
from
nine
residents
of
the
old
Love
Canal
area
in
1978
for
a
variety
of
volatile
organic
compounds,
including
all
three
of
the
brominated
trihalomethanes.
Bromodichloromethane
was
detected
in
the
blood
of
one
individual;
its
concentration
was
14
µ
g/
L
(
ppb).
Dibromochloromethane
and
bromoform
were
not
detected.

Antoine
et
al.
(
1986)
analyzed
the
blood
of
250
environmentally
sensitive
patients
for
18
volatile
organic
compounds.
Bromoform
concentrations
ranged
from
undetectable
to
3.4
µ
g/
L
(
ppb),
with
a
mean
of
0.6
µ
g/
L
(
ppb).

Ashley
et
al.
(
1994)
analyzed
samples
of
whole
blood
of
600
or
more
people
in
the
United
States
who
participated
in
the
Third
National
Health
and
Nutrition
Examination
Survey
(
NHANES
III)
for
32
volatile
organic
compounds
using
analytical
methods
designed
to
measure
extremely
low
concentrations.
Bromodichloromethane,
with
a
detection
limit
of
0.009
µ
g/
L
(
ppb)
was
detected
only
in
14%
of
1072
samples.
Dibromochloromethane,
with
a
detection
limit
of
0.013
µ
g/
L
(
ppb),
was
detected
in
only
12%
of
1035
samples.
Using
unprocessed
commercial
Vacutainer
Tubes,
Ashely
et
al.
(
1994)
initially
obtained
measures
of
bromoform
concentrations
in
blood
similar
to
those
reported
by
Antoine
et
al.
(
1986).
However,
using
Vacutainer
Tubes
that
had
been
processed
to
removed
VOCs
prior
to
use,
Ashely
et
al.
(
1994)
detected
bromoform
in
less
than
10%
of
samples
analyzed
at
a
detection
limit
of
0.027
µ
g/
L
(
ppb).
Wallace
(
1997)
obtained
the
summary
statistics
for
bromodichloromethane
and
dibromochloromethane,
which
were
not
published
in
Ashely
et
al.'
s
(
1994)
paper.
The
mean
(
±
SD)
of
the
measured
blood
concentrations
were
0.0077
±
0.0178
and
0.00886
±
0.00856
µ
g/
L
(
ppb),
respectively.
The
median
values
were
below
the
limit
of
detection.
The
upper
90th
percentile
values
were
0.0122
and
0.0151
µ
g/
L
(
ppb),
respectively.

Weisel
et
al.
(
1999)
measured
brominated
trihalomethanes
in
the
exhaled
breath
of
female
subjects
after
showering.
The
study
authors
recruited
49
women
who
had
previously
participated
in
a
case­
control
study
on
neural
tube
birth
defects
from
locations
throughout
the
state
of
New
Jersey
(
Klotz
and
Pyrch,
1999).
The
method
used
to
select
the
subjects
provided
a
wide
range
of
brominated
trihalomethane
exposures
within
the
home,
in
contrast
to
a
distribution
of
exposures
that
might
exist
within
a
single
water
distribution
system
or
within
the
general
population.
Exposure
to
brominated
trihalomethanes
was
estimated
by
collection
of
duplicate
cold
tap
water
samples,
collection
of
a
15­
minute
air
sample,
and
responses
to
a
48­
hour
recall
questionnaire
on
water
use
in
the
home.
Post­
shower
whole
breath
samples
were
collected
by
having
the
subject
blow
into
a
Tedlar
®
sampling
bag
at
the
conclusion
of
a
shower.
Background
breath
samples
were
collected
at
a
subsequent
home
visit
by
the
investigators.
Valid
samples
were
obtained
from
33
of
the
subjects.
However,
the
time
of
post­
shower
sample
collection
as
reported
by
the
subjects
varied
from
immediately
after
the
shower
to
20
minutes
later.
As
noted
by
the
authors,
the
delay
in
sample
collection
is
an
important
determinant
in
breath
concentrations
because
trihalomethane
breath
concentration
declines
exponentially
after
exposure
ceases.
As
a
result,
Draft
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or
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February
20,
2003
IV
­
33
each
subject
was
assigned
to
one
of
three
groups:
1)
breath
sample
collected
within
5
minutes
after
completion
of
shower
(
Group
A;
n
=
13);
2)
breath
sample
collected
within
5
to
20
minutes
after
completion
of
shower
(
Group
B;
n=
14);
or
3)
breath
sample
collected
more
than
20
minutes
after
showering
(
Group
C;
n=
6).
The
breath
concentrations
on
individual
brominated
trihalomethanes
for
each
group
were
compared
to
measured
water
concentrations
and
estimates
of
exposure
(
calculated
as
the
product
of
the
water
concentration
and
reported
duration
of
the
shower;
shower
duration
data
and
calculated
exposure
estimates
were
not
reported).

The
mean
(
±
standard
deviation)
concentrations
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
were
5.7
±
8.6,
2.0
±
2.1,
and
0.73
±
0.90

g/
L
(
ppb),
respectively.
The
median
values
for
the
three
compounds
were
2.6,
1.4,
and
0.45

g/
L
(
ppb),
respectively.
Bromodichloromethane
showed
significant
correlations
for
breath
and
water
concentration
and
breath
and
shower
exposure
for
Groups
A
and
B.
Significant
correlations
for
dibromochloromethane
and
bromoform
were
found
for
Group
A
participants.
Analytical
variability
related
to
low
concentrations
of
dibromochloromethane
and
bromoform
(
near
the
detection
limit)
may
have
obscured
trends
in
the
data
for
Group
B.
source
of
in
the
houses
and
found
significant
correlations
between
the
water
concentration
of
each
brominated
trihalomethane
and
the
concentration
of
that
trihalomethane
in
expired
air
if
the
air
samples
were
collected
within
5
minutes
of
showering.
Results
of
statistical
analysis
for
Group
C
were
not
reported
because
the
sample
size
was
small
and
the
authors
considered
the
results
questionable.
The
observed
results
were
considered
consistent
with
showering
being
a
source
of
exposure
to
brominated
trihalomethanes.

Backer
et
al.
(
2000)
examined
levels
of
brominated
trihalomethanes
in
whole
blood
following
three
types
of
water
use
events
by
adult
volunteers:
showering
for
10
minutes
in
tap
water
(
n=
11);
bathing
for
10
minutes
in
a
tub
filled
with
tap
water
(
n=
10);
or
consumption
of
one
liter
of
tap
water
over
a
10
minute
period
(
n=
10).
Each
participant
provided
a
blood
sample
immediately
before
exposure,
10
minutes
after
exposure
ended,
and
30
minutes
(
showering
and
bathing)
or
one
hour
(
ingestion)
after
exposure.
Tap
water
and
blood
samples
were
analyzed
by
purge­
and­
trap/
gas
chromatography/
mass
spectrometry
with
detection
capability
in
the
parts
per
quadrillion
range.
Bromoform
was
not
detected
in
either
tap
water
or
whole
blood.
Mean
tap
water
concentrations
of
bromodichloromethane
and
dibromochloromethane
were
6
µ
g/
L
and
1.1
µ
g/
L,
respectively.
The
highest
levels
of
these
compounds
in
whole
blood
occurred
10
minutes
after
exposure
had
ended.
The
second
post­
exposure
measurements
showed
that
blood
levels
of
both
compounds
had
decreased,
but
were
still
above
the
pre­
exposure
baseline
levels
in
subjects
who
took
showers
or
baths.
Measurement
data
are
shown
in
Table
IV­
9
below.

The
study
authors
reported
that
similar
relative
findings
were
obtained
for
dibromochloromethane
(
data
shown
graphically
in
the
study
report).
These
data
indicate
a
dramatic
difference
between
the
whole
blood
levels
resulting
from
ingestion
and
those
resulting
from
bathing
or
showering
(
including
dermal,
inhalation,
and
possibly
ingestion
exposure).
Blood
level
increases
observed
for
each
compound
after
ingestion
of
one
liter
of
water
were
less
than
10%
of
those
observed
after
bathing
or
showering
for
10
minutes.
The
blood
level
increases
observed
for
each
compound
10
minutes
after
bathing
or
showering
were
significantly
Draft
­
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or
Quote
February
20,
2003
IV
­
34
Table
IV­
9
Mean
Bromodichloromethane
Concentrations
in
Blood
Following
Three
Types
of
Water
Use
Events
Water
Use
Event
Median
Bromodichloromethane
Concentration
in
Whole
Blood
(
pg/
mL)

Pre­
exposure
10
minutes
post­
exposure
30
or
60
minutes
postexposure
10
Minute
Shower
3.3
19.4
10.3
(
30
min)

10
Minute
Bath
2.3
17.0
9.9
(
30
min)

Ingestion
of
1
L
2.6
3.8
2.8
(
60
min)

increased
(
p<
0.01)
when
compared
to
the
post­
ingestion
blood
levels.
Measurable
levels
of
bromodichloromethane
and
dibromochloromethane
in
the
pre­
exposure
whole
blood
samples
were
attributed
to
recent
prior
exposure
or
to
bioaccumulation
after
repeated
exposure
to
tap
water.

In
addition
to
the
differences
observed
in
whole
blood
levels
of
bromodichloromethane
and
dibromochloromethane
among
exposure
groups,
the
study
authors
observed
that
the
blood
concentration
data
for
each
chemical
occurred
in
two
clusters
within
each
exposure
group.
The
mean
increases
for
the
two
clusters
observed
after
bathing
or
showering
were
significantly
different
for
bromodichloromethane.
The
same
individuals
who
had
greater
increases
of
bromodichloromethane
also
experienced
greater
increases
of
dibromochloromethane
and
chloroform
in
the
blood
after
bathing
or
showering.
The
underlying
basis
for
the
observed
clustering
is
unknown,
but
was
not
related
to
gender.
The
study
authors
suggested
that
polymorphic
expression
of
a
metabolizing
enzyme
(
e.
g.
glutathione­
S­
transferase
theta)
or
differences
in
fitness
level
(
resulting
in
inhalation
of
larger
volumes
of
air)
may
have
accounted
for
the
observed
pattern.
However,
they
noted
that
differences
in
fitness
level
would
more
likely
be
expected
to
result
in
a
continuous
distribution.

Lynberg
et
al.
(
2001)
conducted
a
field
study
in
Corpus
Christi,
Texas,
and
Cobb
County,
Georgia,
to
evaluate
exposure
measures
for
disinfection
by­
products,
including
brominated
trihalomethanes.
These
areas
were
selected
for
study
based
on
the
following
criteria:
1)
relatively
high
trihalomethane
concentrations
relative
to
national
averages;
2)
high
intrasystem
differences
that
would
result
in
a
potential
exposure
gradient
across
the
study
population;
3)
one
water
distribution
system
with
predominately
chlorinated
species
of
trihalomethanes
(
i.
e.,
chloroform)
and
one
water
system
with
predominately
brominated
trihalomethanes;
and
4)
a
water
utility
service
population
large
enough
to
allow
rapid
selection
of
25
mothers
per
geographic
area
who
had
given
birth
to
healthy
babies
from
June,
1998
through
May,
1999.
Exposure
to
individual
trihalomethanes
was
assessed
by
collection
of
blood
and
water
samples
and
by
collection
of
information
on
water
use
patterns
and
tap
water
characteristics.
Whole
blood
samples
were
collected
before
and
after
showering.
Levels
of
individual
trihalomethanes
were
determined
for
samples
collected
in
the
home
of
participants,
in
the
distribution
system,
and
at
the
water
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
35
treatment
plants.
A
modified
version
of
the
Total
Exposure
Model
(
TEM)
was
used
to
estimate
uptake
of
trihalomethanes
into
the
bloodstream
(
data
for
chloroform
exposure
were
presented
for
one
individual
in
Corpus
Christi).
The
results
of
the
study
indicate
that
concentration
of
individual
trihalomethanes
varied
by
site
and
location
within
the
water
system
(
Table
IV­
10).
In
Corpus
Christi
water
samples,
brominated
trihalomethanes
accounted
for
71%
of
the
total
trihalomethane
concentration
by
weight.
In
contrast,
brominated
trihalomethanes
accounted
for
only
12%
of
the
trihalomethanes
in
Cobb
County
water
samples.
Significant
differences
(
p
=
0.0001)
in
the
blood
levels
of
dibromochloromethane
and
bromoform
were
observed
between
study
locations
(
Table
IV­
11).
The
differences
between
locations
were
evident
both
before
and
after
showering.
The
study
authors
indicated
that
there
was
considerable
variability
in
blood
levels
of
trihalomethanes
among
participants
from
a
single
location.
For
example,
pre­
shower
chloroform
blood
levels
in
Cobb
County
ranged
from
130
ppt
to
1100
ppt.
No
data
were
presented
for
the
brominated
trihalomethanes.
The
variability
was
tentatively
attributed
to
different
patterns
of
household
water
use
among
participants.
Significant
increases
(
p
=
0.0001)
in
blood
levels
of
all
brominated
trihalomethanes
were
observed
after
showering.
The
increases
in
dibromochloromethane
and
bromoform
were
significantly
greater
in
Corpus
Christi
than
in
Cobb
County.
No
TEM
modeling
data
were
presented
for
brominated
trihalomethanes.
However,
TEM
results
presented
for
chloroform
exposure
for
one
study
participant
who
consumed
bottled
water
indicated
that
inhalation
exposure
in
the
household
accounted
for
approximately
98%
of
the
calculated
24­
hour
chloroform
dose,
with
the
remainder
attributed
to
the
dermal
route.
Overall,
this
study
demonstrates
that
blood
levels
of
brominated
trihalomethanes
vary
significantly
across
populations,
with
water
quality
characteristics
and
water
use
activities
being
important
variables.

Table
IV­
10
Median
Tap
Water
Trihalomethane
Levels
(
ppb)
in
Cobb
County
and
Corpus
Christi
Homes,
Water
Treatment
Plants,
and
Distribution
Systems
Trihalomethane
Cobb
County
Corpus
Christi
Home
(
n=
25)
Distribution
System
(
n=
20)
Water
Treatment
Plant
(
n=
7)
Home
(
n=
25)
Distribution
System
(
n=
30)
Water
Treatment
Plant
(
n=
20)

Bromodichloromethane
13.5
12.5
9.5
12.2
8.3
9.5
Dibromochloromethane
1.7
2.4
1.4
13.5
12.6
14.3
Bromoform
NDa
ND
ND
8.7
9.7
11.9
Chloroform
84.8
79
49.5
8.2
4.6
6.7
a
ND,
not
detected
(
detection
limit
<
1ppb)
Draft
­
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Not
Cite
or
Quote
February
20,
2003
IV
­
36
Table
IV­
11
Between
Site
Comparison
of
Median
Blood
Levels
(
ppt)
and
Changes
in
Blood
Levels
(
ppt)
after
Showering
Trihalomethane
Before
Shower
After
Shower
Change
in
Blood
Level
after
Showering
Cobba
Corpusb
Cobb
Corpus
Cobb
Corpus
Bromodichloromethane
6.2
6.8
38
43
30
34
Dibromochloromethane
1.2
7.0
6.1
41
5.0
35
Bromoform
0.3
3.5
0.5
17
0.2
12
Chloroform
70
25
280
57
189
25
a
Cobb
County,
Georgia
b
Corpus
Christi,
TX
2.
Mother's
Milk
Pellizzari
et
al.
(
1982)
analyzed
the
milk
of
eight
nursing
mothers
for
various
compounds,
including
bromodichloromethane
and
dibromochloromethane.
The
samples
were
collected
from
49
lactating
women
living
in
the
vicinity
of
chemical
manufacturing
plants
and/
or
industrial
user
facilities
in
Bridgeville,
PA,
Bayonne,
NJ,
Jersey
City,
NJ,
and
Baton
Rouge
LA.
Both
compounds
were
identified
in
one
of
the
eight
samples.
Actual
concentrations
and
detection
limits
were
not
reported.
Kroneld
and
Reunanen
(
1990)
did
not
detect
any
of
the
brominated
trihalomethanes
in
human
milk
in
a
study
conducted
in
Turku,
Finland.

E.
Summary
Brominated
trihalomethanes
are
found
in
virtually
all
water
treated
for
drinking;
however,
concentrations
of
individual
forms
vary
widely
depending
on
the
type
of
water
treatment,
locale,
time
of
year,
sampling
point
in
the
distribution
system,
and
source
of
the
drinking
water.
Occurrence
data
for
brominated
trihalomethanes
are
available
from
13
national
surveys
and
9
additional
studies
that
are
more
restricted
in
scope.
The
procedures
used
for
sampling
processing
and
storage
and
calculation
of
summary
statistics
should
be
carefully
considered
when
evaluating
and
comparing
brominated
trihalomethane
occurrence
data.
Some
methods
restrict
trihalomethane
formation
by
refrigeration
or
the
use
of
quenching
agents,
whereas
others
maximize
trihalomethane
formation
by
storage
at
room
temperature.
Approaches
to
data
summarization
vary
by
study
in
the
treatment
of
data
below
the
analytical
detection
level
or
minimum
reporting
level.

When
all
available
national
survey
data
are
considered,
bromodichloromethane
concentrations
in
drinking
water
range
from
below
the
detection
limit
to
183
µ
g/
L
(
ppb),
while
dibromochloromethane
and
bromoform
concentrations
range
from
below
the
detection
limit
to
280
µ
g/
L
(
ppb).
When
data
for
the
three
brominated
trihalomethanes
are
compared,
the
frequency
of
Draft
­
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or
Quote
February
20,
2003
IV
­
37
detection
and
measured
concentrations
of
bromodichloromethane
in
drinking
water
supplies
tend
to
be
higher
than
those
for
dibromochloromethane.
Bromoform
is
detected
less
frequently
and
at
lower
concentrations
than
the
other
two
brominated
trihalomethanes,
except
in
some
ground
waters.
Concentrations
of
all
trihalomethanes
in
drinking
water
were
generally
lower
when
the
raw
water
is
obtained
from
ground
water
sources
rather
than
surface
water
sources.
The
most
recent
national
survey
data
are
those
collected
by
the
U.
S.
EPA
under
the
Information
Collection
Rule
(
ICR).
Monitoring
data
were
collected
over
an
18­
month
period
between
July
1997
and
December
1998
from
approximately
300
water
systems
operating
501
plants
and
serving
at
least
100,000
people.
Summary
occurrence
data
stratified
by
raw
water
source
(
groundwater
or
surface
water)
are
available
for
finished
water,
the
distribution
system
(
DS)
average,
and
the
DS
high
values.
The
mean,
median,
and
90th
percentile
values
for
surface
water
DS
average
concentrations
in
the
ICR
survey
are
8.6,
70.2,
and
20.3
µ
g/
L,
respectively,
for
bromodichloromethane
(
range
of
individual
values
0
­
65.8
µ
g/
L);
2.4,
4.72,
and
13.2
µ
g/
L,
respectively,
for
dibromochloromethane
(
range
0
­
67.3);
and
0.
1.18,
and
3.10,
respectively,
for
bromoform
(
range
0
­
3.43).

Relatively
few
studies
have
analyzed
non­
beverage
foods
for
the
occurrence
of
brominated
trihalomethanes.
In
the
few
studies
available,
bromodichloromethane
has
been
detected
in
nonbeverage
foods
(
i.
e.,
in
one
sample
of
butter
at
7
ppb,
in
three
samples
of
ice­
cream
at
0.6
to
2.3
ppb,
in
6
of
10
samples
of
bean
curd
at
1.2
to
5.2
ppb,
and
in
one
sample
of
bacon
(
probably
below
the
minimal
quantitation
limit)).
In
addition,
bromodichloromethane
was
detected
in
one
sample
each
of
eleven
foods
out
of
70
tested
in
14
Market
Baskets
for
the
FDA
Total
Diet
Study.
The
detected
concentrations
ranged
from
10
to
37
ppb
for
individual
food
items.
Studies
that
analyzed
non­
beverage
foods
for
dibromochloromethane
and
bromoform
detected
neither
compound
in
any
of
the
samples.
Brominated
trihalomethanes
have
been
detected
in
up
to
a
third
or
one
half
of
the
types
of
prepared
beverages
examined
in
some
studies,
being
detected
most
frequently
in
colas
and
other
carbonated
soft
drinks.
Bromodichloromethane
has
been
found
most
frequently
of
the
three
compounds
and
bromoform
the
least
frequently.
Bromodichloromethane
was
detected
in
approximately
half
of
the
prepared
beverages
examined
by
McNeal
et
al.
(
1995)
in
the
United
States
and
in
all
of
13
soft
drinks
that
they
analyzed.
With
the
exception
of
one
of
the
13
soft
drinks
examined
by
McNeal
et
al.
(
1995)
with
a
concentration
of
12
ppb,
none
of
the
at
least
18
other
measured
concentrations
of
bromodichloromethane
in
soft
drinks
described
above
(
three
from
Entz
et
al.
(
1982),
three
from
Uhler
and
Diachenko
(
1987),
and
the
remaining
12
from
McNeal
et
al.
(
1995))
exceeded
a
value
of
4
ppb.
Bromodichloromethane
was
detected
in
one
sample
of
fruit
juice
at
5
ppb.

Exposure
to
brominated
trihalomethanes
via
ingestion
of
drinking
water
was
estimated
using
data
obtained
for
disinfectants
and
disinfection
byproducts
under
the
Information
Collection
Rule
(
ICR).
ICR
data
offer
several
advantages
over
other
national
studies
for
purposes
of
estimating
national
exposure
levels
of
adults
in
the
United
States
to
brominated
trihalomethanes
via
ingestion
of
drinking
water.
First,
they
are
recent
and
reflect
relatively
current
conditions.
Second,
data
of
very
similar
quality
and
quantity
were
collected
systematically
from
a
large
number
of
plants
(
501)
and
systems
(
approximately
300),
including
both
surface
and
ground
water
systems.
Third,
the
mean,
median,
and
90th
percentile
value
were
estimated
on
the
basis
of
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
38
all
samples
taken,
not
just
the
sample
detects.
Thus,
these
descriptive
statistics
are
representative
of
the
exposures
of
the
entire
populations
served
by
those
systems,
not
just
the
populations
served
by
systems
with
higher
concentrations
of
these
compounds.
However,
this
study
can
not
be
considered
representative
of
smaller
public
water
supplies
or
water
supplies
from
the
most
highly
industrialized
or
contaminated
areas.

Exposure
was
calculated
by
multiplying
the
concentration
of
individual
brominated
trihalomethanes
in
drinking
water
by
the
average
daily
intake,
assuming
that
each
individual
consumes
two
liters
of
water
per
day.
The
annual
median,
mean,
and
upper
90th
percentile
values
are
presented
for
both
surface
and
ground
water
systems.
Assuming
that
the
DS
High
value
actually
represents
the
average
exposure
level
of
persons
served
by
one
plant
distribution
pipe
with
the
longest
water­
residence
time,
the
DS
High
value
might
be
used
to
estimate
a
high­
end
exposure
level.

For
bromodichloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
17,
20,
and
40
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromodichloromethane
from
ground
water
systems
are
lower
 
3.6,
8.1,
and
22
µ
g/
person/
day,
respectively.
For
dibromochloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
4.8,
9.4,
and
26
µ
g/
person/
day,
respectively.
The
corresponding
values
for
populations
exposed
to
dibromochloromethane
from
groundwater
system
are
lower
 
2.7,
6.2,
and
18
µ
g/
person/
day,
respectively.
For
bromoform,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
near
0,
2.4,
and
6.2
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromoform
from
ground
water
systems
are
higher
 
0.65,
3.8,
and
9.6
µ
g/
person/
day,
respectively.

For
purposes
of
comparison,
estimates
of
ingestion
exposure
to
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
drinking
water
were
also
estimated
from
data
collected
in
other,
older
studies.
Ingestion
from
ground
water
supplies
was
estimated
from
the
median
levels
found
in
the
Ground
Water
Supply
Survey
conducted
by
U.
S.
EPA
in
1980­
81.
Based
on
the
range
of
median
levels
(
1.4
 
2.1
µ
g/
L
(
ppb))
and
a
consumption
rate
of
two
liters
per
day,
the
median
ingestion
exposure
to
bromodichloromethane
may
range
from
2.8
to
4.2
µ
g/
day.
Similarly,
median
exposure
to
dibromochloromethane
may
range
from
4.2
to
7.8
µ
g/
day,
and
for
bromoform,
median
exposure
may
range
from
4.8
to
8.4
µ
g/
day.
Exposure
to
bromodichloromethane
from
surface
water
supplies
can
be
estimated
based
on
the
range
of
median
values
observed
under
different
conditions
in
the
National
Organics
Monitoring
Survey
conducted
by
U.
S.
EPA
in
1976­
1977,
which
mainly
sampled
surface
water
systems.
Based
on
a
range
of
5.9
 
14
µ
g/
L
(
ppb),
exposure
to
bromodichloromethane
from
surface
water
is
estimated
to
be
between
12
and
28
µ
g/
day.
Similarly,
based
on
the
range
of
medians
reported
for
dibromochloromethane
concentrations,
the
median
exposure
is
estimated
to
be
up
to
6
µ
g/
day.
The
median
levels
of
bromoform
in
the
surface
water
supplies
have
been
found
to
be
less
than
the
EPA
Drinking
Water
minimum
reporting
levels
(
MRLs)
of
0.5
 
1
µ
g/
L
(
ppb).
An
estimate
of
exposure
based
on
the
MRLs
will
be
overly
conservative
because
the
actual
concentration
of
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
39
bromoform
is
not
detectable.
Based
on
the
range
of
MRLs,
0.5
 
1
µ
g/
L
(
ppb),
the
exposure
to
bromoform
is
estimated
to
range
from
1
to
2
µ
g/
day
for
surface
water
supplies.

Ingestion
exposure
to
brominated
trihalomethanes
in
drinking
water
can
also
be
estimated
from
the
concentrations
found
at
the
tap
in
the
U.
S.
EPA's
Total
Exposure
Assessment
Methodology
(
TEAM)
study.
Estimates
of
the
average
of
the
population
intakes
for
ingestion
of
bromodichloromethane
from
drinking
water
range
from
0.42
to
42

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
2.0
to
90

g/
person/
day.
Estimates
of
the
average
population
intake
of
dibromochloromethane
from
drinking
water
range
from
0.2
to
56

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
0.9
to
86

g/
person/
day.
Estimates
of
the
average
of
the
population
intakes
of
bromoform,
for
those
areas
in
which
bromoform
was
measurable
in
a
majority
of
the
samples,
range
from
1.6
to
16.2

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
2.4
to
26

g/
person/
day.
Four
of
the
six
locations
in
the
TEAM
study,
however,
had
a
low
frequency
(
less
than
10%)
of
detection
of
bromoform
in
measurable
quantities.

Sources
of
uncertainty
in
these
estimates
of
ingestion
exposure
include
use
of
different
analytical
methods,
failure
to
report
quantitation
limits,
using
measures
near
the
detection
limit,
failure
to
report
how
nondetects
are
handled
when
averaging
values
(
e.
g.,
set
to
zero
or
one
half
the
detection
limit),
and
failure
to
report
sample
storage
method
and
duration.
In
addition,
many
environmental
factors
influence
the
concentrations
of
these
compounds
in
drinking
water
at
the
tap
and
in
vended
or
bottled
waters
used
for
drinking.
These
factors
include
season
and
temperature,
geographic
location,
source
of
water,
residence
time
in
distribution
system,
and
others.

Average
daily
intake
of
dibromochloromethane
via
ingestion,
dermal
contact,
and
inhalation
of
compound
volatilized
during
household
use
were
also
estimated
for
determination
of
the
Relative
Source
Concentration
(
RSC).
Intake
for
ingestion
was
calculated
using
mean
intake
rates
of
1.2
or
0.6
L/
day
for
total
and
direct
intake
(
NRC,
1999),
respectively.
Direct
intake
includes
consumption
of
water
directly
from
the
tap,
but
does
not
include
intake
of
tap
water
used
for
preparation
of
heated
items
such
tea,
coffee,
or
soup.
Based
on
the
ICR
distribution
system
average
concentration
of
4.72
µ
g/
L
for
dibromochloromethane
in
surface
water,
the
average
daily
total
and
direct
and
ingestion
intakes
would
be
5.7
and
2.8
µ
g/
day,
respectively.
The
average
dermal
uptake
of
dibromochloromethane
was
estimated
to
be
2
µ
g
per
shower
or
bathing
event.
Average
daily
intake
via
inhalation
of
dibromochloromethane
volatilized
during
to
be
7
µ
g/
day
for
the
volatilized
compound.
Parallel
calculations
were
not
performed
for
bromodichloromethane
or
bromoform,
because
these
compounds
are
probable
carcinogens.
Therefore,
in
accordance
with
U.
S.
EPA
policy,
RSC
analysis
was
not
conducted.

Some
data
on
the
occurrence
of
brominated
trihalomethanes
in
foods
and
beverages
are
available
from
studies
conducted
in
Italy,
Japan,
and
Finland.
These
studies
were
also
limited
in
scope
to
examination
of
relatively
few
food
or
beverage
items.
Bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
measured
in
foods
and
beverages
in
Italy,
Japan
and
Finland
ranged
from
undetectable
to
40
ppb,
undetectable
to
13.9
ppb,
and
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
40
undetectable
to
10.7
ppb,
respectively.
Because
of
possible
differences
in
water
disinfection
or
food
processing
practices,
these
data
may
not
be
representative
of
concentrations
in
foods
and
beverages
produced
in
the
U.
S.

Concentrations
in
outdoor
air
were
variable
from
site
to
site.
When
data
from
several
urban/
suburban
and
source­
dominated
sites
in
Texas,
Louisiana,
North
Carolina
and/
or
Arkansas
were
combined,
the
resulting
average
outdoor
air
concentrations
were
110
ppt
(
0.74
µ
g/
m3)
for
bromodichloromethane,
3.8
ppt
(
0.032
µ
g/
m3)
for
dibromochloromethane,
and
3.6
ppt
(
0.037
µ
g/
m3)
for
bromoform.
A
regional
study
conducted
at
several
sites
in
southern
California
found
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
35%,
17%,
and
31%
of
the
samples,
respectively.
The
maximum
concentrations
observed
were
40
ppt
(
0.27
µ
g/
m3)
for
bromodichloromethane;
290
ppt
(
2.5
µ
g/
m3)
for
dibromochloromethane;
310
ppt
(
3.2
µ
g/
m3)
for
bromoform.
Bromodichloromethane
was
detected
in
64%
(
n=
11)
and
17%
(
n=
6)
of
personal
air
samples
collected
in
Texas
and
North
Carolina.
The
detected
concentrations
ranged
from
0.12
to
4.36
µ
g/
m3
(
0.017
to
0.65
ppb).
Dibromochloromethane
was
not
detected.

Mean
concentrations
in
indoor
air
ranged
from
0.38
to
0.75
µ
g/
m3
for
bromodichloromethane
0.44
to
0.53
µ
g/
m3
for
dibromochloromethane,
and
0.29
to
0.35
µ
g/
m3
for
bromoform,
as
determined
from
15
minute
samples
collected
in
48
New
Jersey
residences.
In
a
separate
study,
levels
of
brominated
trihalomethanes
in
indoor
air
were
locally
increased
(
e.
g.,
in
shower/
bath
enclosures
and
vanity
areas)
during
showering
and
bathing
events.
The
levels
of
individual
brominated
trihalomethanes
in
air
were
reported
to
be
consistent
with
the
levels
in
tap
water.

The
use
of
chlorine
to
disinfect
swimming
pools
and
hot
tubs
results
in
the
formation
of
brominated
trihalomethanes.
Swimming
pool
and
hot
tub
users
are
potentially
exposed
to
brominated
trihalomethanes
via
dermal
contact,
ingestion,
and
inhalation
of
compounds
released
to
the
overlying
air.
As
a
result,
swimming
pool
and
hot
tub
users
may
experience
greater
overall
exposures
to
brominated
trihalomethanes
than
the
general
population.
One
study
indicated
that
bromodichloromethane,
dibromochloromethane,
and
bromoform
concentrations
in
swimming
pool
and
hot
tub
water
ranged
from
1
to
105
µ
g/
L
(
ppb),
from
0.1
to
48
µ
g/
L
(
ppb),
and
from
less
than
0.1
to
62
µ
g/
L
(
ppb),
respectively.
Concentrations
of
the
same
brominated
trihalomethanes
in
the
air
two
meters
above
the
pool
water
ranged
from
less
than
0.1
to
14
µ
g/
m3
(
0.015
 
2.09
ppb),
from
less
than
0.1
 
10
µ
g/
m3
(
0.011
 
1.2
ppb),
and
from
less
than
0.1
to
5.0
µ
g/
m3
(
0.0097
 
0.48
ppb),
respectively.
Data
from
several
studies
confirm
the
uptake
of
brominated
trihalomethanes
from
swimming
pools
and
environs
by
dermal
and/
or
inhalation
pathways.

No
data
for
occurrence
of
brominated
trihalomethanes
in
soil
were
available
in
the
materials
reviewed
for
this
document.
The
chemical
and
physical
properties
of
the
brominated
trihalomethanes
indicate
that
they
should
volatilize
readily
from
wet
or
dry
soil
surfaces.
Therefore,
ingestion
of
soil
is
not
expected
to
be
a
significant
route
of
exposure.

Exposure
to
brominated
trihalomethanes
via
ingestion
of
drinking
water
was
estimated
using
data
obtained
for
disinfectants
and
disinfection
byproducts
under
the
Information
Collection
Rule
(
ICR).
ICR
data
offer
several
advantages
over
other
national
studies
for
purposes
of
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
IV
­
41
estimating
national
exposure
levels
of
adults
in
the
United
States
to
brominated
trihalomethanes
via
ingestion
of
drinking
water.
First,
they
are
recent
and
reflect
relatively
current
conditions.
Second,
data
of
very
similar
quality
and
quantity
were
collected
systematically
from
a
large
number
of
plants
(
501)
and
systems
(
approximately
300),
including
both
surface
and
ground
water
systems.
Third,
the
mean,
median,
and
90th
percentile
value
were
estimated
on
the
basis
of
all
samples
taken,
not
just
the
sample
detects.
Thus,
these
descriptive
statistics
are
representative
of
the
exposures
of
the
entire
populations
served
by
those
systems,
not
just
the
populations
served
by
systems
with
higher
concentrations
of
these
compounds.
However,
this
study
can
not
be
considered
representative
of
smaller
public
water
supplies
or
water
supplies
from
the
most
highly
industrialized
or
contaminated
areas.

Exposure
was
calculated
by
multiplying
the
concentration
of
individual
brominated
trihalomethanes
in
drinking
water
by
the
average
daily
intake,
assuming
that
each
individual
consumes
two
liters
of
water
per
day.
The
annual
median,
mean,
and
upper
90th
percentile
values
are
presented
for
both
surface
and
ground
water
systems.
Assuming
that
the
DS
High
value
actually
represents
the
average
exposure
level
of
persons
served
by
one
plant
distribution
pipe
with
the
longest
water­
residence
time,
the
DS
High
value
might
be
used
to
estimate
a
high­
end
exposure
level.

For
bromodichloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
17,
20,
and
40
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromodichloromethane
from
ground
water
systems
are
lower
 
3.6,
8.1,
and
22
µ
g/
person/
day,
respectively.
For
dibromochloromethane,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
4.8,
9.4,
and
26
µ
g/
person/
day,
respectively.
The
corresponding
values
for
populations
exposed
to
dibromochloromethane
from
groundwater
system
are
lower
 
2.7,
6.2,
and
18
µ
g/
person/
day,
respectively.
For
bromoform,
the
median,
mean,
and
90th
percentile
population
exposures
from
surface
water
systems
are
estimated
to
be
near
0,
2.4,
and
6.2
µ
g/
person/
day,
respectively.
The
same
values
for
populations
exposed
to
bromoform
from
ground
water
systems
are
higher
 
0.65,
3.8,
and
9.6
µ
g/
person/
day,
respectively.

For
purposes
of
comparison,
estimates
of
ingestion
exposure
to
bromodichloromethane,
dibromochloromethane,
and
bromoform
in
drinking
water
were
also
estimated
from
data
collected
in
other,
older
studies.
Ingestion
from
ground
water
supplies
was
estimated
from
the
median
levels
found
in
the
Ground
Water
Supply
Survey
conducted
by
U.
S.
EPA
in
1980­
81.
Based
on
the
range
of
median
levels
(
1.4
 
2.1
µ
g/
L
(
ppb))
and
a
consumption
rate
of
two
liters
per
day,
the
median
ingestion
exposure
to
bromodichloromethane
may
range
from
2.8
to
4.2
µ
g/
day.
Similarly,
median
exposure
to
dibromochloromethane
may
range
from
4.2
to
7.8
µ
g/
day,
and
for
bromoform,
median
exposure
may
range
from
4.8
to
8.4
µ
g/
day.
Exposure
to
bromodichloromethane
from
surface
water
supplies
can
be
estimated
based
on
the
range
of
median
values
observed
under
different
conditions
in
the
National
Organics
Monitoring
Survey
conducted
by
U.
S.
EPA
in
1976­
1977,
which
mainly
sampled
surface
water
systems.
Based
on
a
range
of
5.9
 
14
µ
g/
L
(
ppb),
exposure
to
bromodichloromethane
from
surface
water
is
estimated
to
be
between
12
and
28
µ
g/
day.
Similarly,
based
on
the
range
of
medians
reported
for
Draft
­
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or
Quote
February
20,
2003
IV
­
42
dibromochloromethane
concentrations,
the
median
exposure
is
estimated
to
be
up
to
6
µ
g/
day.
The
median
levels
of
bromoform
in
the
surface
water
supplies
have
been
found
to
be
less
than
the
EPA
Drinking
Water
minimum
reporting
levels
(
MRLs)
of
0.5
 
1
µ
g/
L
(
ppb).
An
estimate
of
exposure
based
on
the
MRLs
will
be
overly
conservative
because
the
actual
concentration
of
bromoform
is
not
detectable.
Based
on
the
range
of
MRLs,
0.5
 
1
µ
g/
L
(
ppb),
the
exposure
to
bromoform
is
estimated
to
range
from
1
to
2
µ
g/
day
for
surface
water
supplies.

Ingestion
exposure
to
brominated
trihalomethanes
in
drinking
water
can
also
be
estimated
from
the
concentrations
found
at
the
tap
in
the
U.
S.
EPA's
Total
Exposure
Assessment
Methodology
(
TEAM)
study.
Estimates
of
the
average
of
the
population
intakes
for
ingestion
of
bromodichloromethane
from
drinking
water
range
from
0.42
to
42

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
2.0
to
90

g/
person/
day.
Estimates
of
the
average
population
intake
of
dibromochloromethane
from
drinking
water
range
from
0.2
to
56

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
<
0.9
to
86

g/
person/
day.
Estimates
of
the
average
of
the
population
intakes
of
bromoform,
for
those
areas
in
which
bromoform
was
measurable
in
a
majority
of
the
samples,
range
from
1.6
to
16.2

g/
person/
day.
The
upper
90th
percentile
estimates
range
from
2.4
to
26

g/
person/
day.
Four
of
the
six
locations
in
the
TEAM
study,
however,
had
a
low
frequency
(
less
than
10%)
of
detection
of
bromoform
in
measurable
quantities.

Sources
of
uncertainty
in
these
estimates
of
ingestion
exposure
include
use
of
different
analytical
methods,
failure
to
report
quantitation
limits,
using
measures
near
the
detection
limit,
failure
to
report
how
nondetects
are
handled
when
averaging
values
(
e.
g.,
set
to
zero
or
one
half
the
detection
limit),
and
failure
to
report
sample
storage
method
and
duration.
In
addition,
many
environmental
factors
influence
the
concentrations
of
these
compounds
in
drinking
water
at
the
tap
and
in
vended
or
bottled
waters
used
for
drinking.
These
factors
include
season
and
temperature,
geographic
location,
source
of
water,
residence
time
in
distribution
system,
and
others.

The
RSC
(
relative
source
contribution)
is
the
percentage
of
total
daily
exposure
that
is
attributable
to
tap
water
when
all
potential
sources
are
considered
(
e.
g.,
air,
food,
soil,
and
water).
Ideally,
the
RSC
is
determined
quantitatively
using
nationwide,
central
tendency
and/
or
high­
end
estimates
of
exposure
from
each
relevant
medium.
In
the
absence
of
such
data,
a
default
RSC
ranging
from
20%
to
80%
may
be
used.

The
RSC
used
in
the
current
and
previous
drinking
water
regulations
for
dibromochloromethane
is
80%.
This
value
was
established
by
use
of
a
screening
level
approach
to
estimate
and
compare
exposure
to
dibromochloromethane
from
various
sources.
Information
considered
for
during
this
process
is
summarized
in
Appendix
C.
There
are
some
uncertainties
in
the
80%
RSC
that
are
related
to
the
availability
of
adequate
concentration
data
for
dibromochloromethane
in
media
other
than
water.
Parallel
RSC
calculations
were
not
performed
for
bromodichloromethane
and
bromoform.
The
EPA
has
set
the
regulatory
level
for
these
chemicals
in
drinking
water
at
zero
because
it
has
been
determined
that
they
are
probable
human
Draft
­
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or
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February
20,
2003
IV
­
43
carcinogens.
Therefore,
determination
of
an
RSC
is
not
relevant
for
these
chemicals
because
it
is
the
Agency's
policy
to
perform
RSC
analysis
only
for
noncarcinogens.

Brominated
trihalomethanes
have
been
detected
in
the
blood
and
breast
milk
of
humans.
A
national
survey
of
volatile
organic
compounds
in
whole
blood
detected
bromodichloromethane
dibromochloromethane,
and
bromoform
in
14%,
12%,
and
less
than
10%
of
samples,
respectively,
when
highly
sensitive
analytical
methods
were
applied.
Several
studies
have
demonstrated
that
the
level
of
individual
brominated
trihalomethanes
in
blood
or
breath
increases
shortly
after
exposure
to
these
compounds
in
tap
water
during
bathing
and
showering.
Exposure
during
these
events
may
occur
by
ingestion,
dermal
contact
and/
or
inhalation
of
the
volatilized
compound.
In
studies
which
examined
households
with
differing
concentrations
of
brominated
trihalomethanes
in
tap
water,
the
levels
of
individual
brominated
trihalomethanes
in
blood
or
exhaled
breath
paralleled
the
tap
water
concentration.
The
studies
of
showering
and
bathing
indicate
that
water
use
patterns
and
water
quality
characteristics
are
important
variables
in
determining
the
blood
levels
of
brominated
trihalomethanes.
Dibromochloromethane
was
detected
in
one
of
eight
samples
of
breast
milk
collected
from
women
living
in
the
vicinity
of
U.
S.
chemical
manufacturing
plants
or
user
facilities.
Draft
­
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or
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February
20,
2003
V
­
1
V.
HEALTH
EFFECTS
IN
ANIMALS
A.
Acute
Exposures
This
section
presents
data
on
the
acute
effects
of
brominated
trihalomethanes.
Acute
lethality
values
for
the
brominated
trihalomethanes
are
summarized
in
Table
V­
1.
Additional
acute
toxicity
data
are
summarized
in
Table
V­
2.

1.
Bromodichloromethane
Acute
lethality
of
bromodichloromethane
has
been
investigated
in
mice
and
rats.
LD
50
values
for
male
and
female
ICR
Swiss
mice
were
450
and
900
mg/
kg,
respectively
(
Bowman
et
al.,
1978).
Chu
et
al.
(
1980)
determined
LD
50
values
of
916
and
969
mg/
kg
for
male
and
female
Sprague­
Dawley
rats,
respectively.

Bowman
et
al.
(
1978)
administered
bromodichloromethane
in
a
single
gavage
dose
in
Emulphor
®
:
alcohol:
saline
(
1:
1:
8)
to
ICR
Swiss
mice
(
10/
sex/
group).
The
administered
doses
ranged
from
500
to
4,000
mg/
kg
(
individual
doses
not
reported).
Sedation
and
anesthesia
occurred
at
500
mg/
kg.
Males
were
more
sensitive
to
the
lethal
effects
of
bromodichloromethane
than
females.

Table
V­
1
Summary
of
LD50
Values
for
Brominated
Trihalomethanes
Compound
LD50
Values
(
mg/
kg)

ICR
Swiss
Mouse
a
Sprague­
Dawley
Rat
b
Male
Female
Male
Female
Bromodichloromethane
450
900
916
969
Dibromochloromethane
800
1200
1186
848
Bromoform
1400
1550
1388
1147
aBowman
et
al.
(
1978)
bChu
et
al.
(
1980)
Draft
­
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or
Quote
February
20,
2003
V
­
2
Table
V­
2
Summary
of
Acute
Toxicity
Studies
for
Brominated
Trihalomethanes
Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Bromodichloromethane
Bowman
et
al.
(
1978)
Mouse
ICR
Swiss
Gavage
(
aqueous)
M,
F
10
Single
dose
500
­
4000
Anesthesia,
sedation
at
500
mg/
kg
NTP
(
1987)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
5
Single
dose
150
300
600
1,250
2,500
Increased
mortality
at
600
mg/
kg­
day.
Lethargy,
labored
breathing
at
1250
mg/
kg­
day
and
above,
100%
mortality
at
the
two
highest
dose
groups
NTP
(
1987)
Mouse
B6C3F1
Gavage
(
oil)
M,
F
5
Single
dose
150
300
600
1,250
2,500
100%
mortality
at
the
two
highest
dose
groups
Lilly
et
al.
(
1994)
Rat
F344
Gavage
(
corn
oil)
(
aqueous)
M
?
Single
dose
0
200
(
LOAEL)
400
Renal
tubule
degeneration
and
necrosis;
alteration
in
markers
of
renal
function
Lilly
et
al.
(
1996)
Rat
F344
Gavage
(
corn
oil
or
water)
M
6
Single
dose
0
200
(
LOAEL)
400
Renal
tubule
necrosis;
alteration
in
markers
of
renal
function
Lilly
et
al.
(
1997)
Rat
F344
Gavage
(
aqueous)
M
5
Single
dose
0
123
164
(
NOAEL)
246
(
LOAEL)
328
492
Decreased
body
weight;
elevated
liver
and
renal
markers
Keegan
et
al.
(
1998)
Rat
F344
Gavage
(
aqueous)
M
6
Single
dose
0
21
31
41
(
NOAEL)
82
(
LOAEL)
123
164
246
Elevated
renal
markers;
decreased
liver
weight
and
body
weight
Dibromochloromethane
Bowman
et
al.
(
1978)
Mouse
ICR
Swiss
Gavage
(
aqueous)
M,
F
10
Single
dose
500
­
4000
Anesthesia,
sedation
Table
V­
2
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
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or
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February
20,
2003
V
­
3
NTP
(
1985)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
5
Single
dose
160
310
630
1250
2,500
Increased
mortality
at
630
mg/
kg
and
above,
with
100%
mortality
in
high
dose
group
NTP
(
1985)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
5
Single
dose
160
310
630
1250
2500
Increased
mortality
in
females
at
630
mg/
kg­
day
and
above
and
in
males
at
310
mg/
kg­
day
and
above;
100%
mortality
in
males
in
two
highest
dose
groups
and
females
in
two
highest
dose
groups
Müller
et
al.
(
1997)
Rat
Wistar
Gavage
(
olive
oil)
M
6
Single
dose
0
83
167
333
667
Transient
decrease
in
blood
pressure
and
heart
rate;
decreased
activity;
effects
on
heart
muscle
contractility
and
changes
in
some
cardiac
parameters
Bromoform
Bowman
et
al.
(
1978)
Mouse
ICR
Swiss
Gavage
(
aqueous)
M,
F
10
Single
dose
500
­
4000
Ataxia,
sedation,
and
anesthesia
at
500
mg/
kg
Chu
et
al.
(
1980)
Rat
SD*
Gavage
(
corn
oil)
M,
F
10
Single
dose
546
765
1071
1500
2100
Sedation,
ataxia,
liver
and
kidney
congestion
NTP
(
1989a)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
5
Single
dose
125
250
500
1,000
2000
No
deaths
at
500
and
lower;
60%
mortality
at
1,000;
100%
mortality
at
2,000;
shallow
breathing
in
two
highest
dose
groups
NTP
(
1989a)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
5
Single
dose
125
250
500
1,000
2000
10%
mortality
at
500
mg/
kg­
day
*
SD,
Sprague­
Dawley
Draft
­
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February
20,
2003
V
­
4
NTP
(
1987)
administered
single
gavage
doses
of
bromodichloromethane
in
corn
oil
to
male
and
female
F344/
N
rats
and
B6C3F
1
mice
(
5/
sex/
dose)
at
150,
300,
600,
1,250,
or
2,500
mg/
kg.
Animals
were
observed
for
14
days,
and
a
necropsy
was
performed
on
at
least
one
male
and
one
female
in
each
dose
group.
All
animals
dosed
with
1,250
or
2,500
mg/
kg
died
before
the
end
of
the
study.
At
600
mg/
kg,
deaths
occurred
in
two
of
five
male
rats,
one
of
five
female
rats,
five
of
five
male
mice,
and
two
of
five
female
mice.
Clinical
signs
observed
in
rats
at
1,250
or
2,500
mg/
kg
included
lethargy
and
labored
breathing.
Clinical
signs
observed
in
mice
at
or
above
600
mg/
kg
included
lethargy,
with
the
exception
that
this
sign
was
not
observed
in
highdose
male
mice.
At
necropsy,
the
liver
from
animals
dosed
with
1,250
or
2,500
mg/
kg
appeared
pale.
No
dose­
related
effects
were
seen
on
body
weight
gain
in
animals
that
survived.

Lilly
et
al.
(
1994)
examined
the
effect
of
vehicle
on
the
toxicity
of
bromodichloromethane
Male
F344
rats
(
number
unknown)
were
administered
a
single
dose
of
0,
200,
or
400
mg/
kg
bromodichloromethane
by
gavage
in
corn
oil
or
in
an
aqueous
10%
Emulphor
®
solution.
Body
weights
were
significantly
decreased
at
400
mg/
kg
only
in
the
animals
receiving
the
aqueous
gavage.
Absolute
and
relative
kidney
weights
were
significantly
increased
at
400
mg/
kg
in
both
vehicles,
with
a
significantly
greater
increase
observed
in
the
animals
gavaged
with
oil
compared
to
those
gavaged
with
10%
Emulphor
®
solution.
Serum
markers
of
hepatotoxicity
were
significantly
increased
at
400
mg/
kg
in
both
vehicles
with
one
nonsignificant
increase
in
the
aqueous
vehicle.
The
increases
were
significantly
greater
for
two
of
these
markers
in
animals
receiving
the
oil
vehicle
compared
to
those
receiving
the
aqueous
vehicle.
Clinical
observations
were
supported
by
histopathology
findings.
Hepatocellular
degeneration
and
necrosis
were
observed
at
400
mg/
kg
in
both
vehicles.
The
difference
in
vehicles
was
reflected
in
more
severe
hepatocellular
degeneration
and
a
higher
incidence
of
centrilobular
necrosis
in
animals
receiving
the
oil
gavage
compared
to
those
receiving
the
aqueous
gavage.
Numerous
increases
in
urinary
markers
of
renal
toxicity
were
observed
24
hours
after
dosing.
Based
on
the
differences
observed,
renal
toxicity
at
200
mg/
kg
was
similar
or
greater
in
the
aqueous
vehicle.
Renal
toxicity
at
400
mg/
kg,
however,
was
greater
in
the
oil
vehicle.
The
time
to
peak
toxicity
was
both
doseand
vehicle­
dependent.
At
200
mg/
kg,
peak
damage
was
observed
at
24
hours
in
both
vehicles.
At
400
mg/
kg,
peak
damage
was
observed
at
48
hours
following
oil
gavage
and
at
24­
36
hours
following
aqueous
gavage.
Histopathology
revealed
both
renal
tubule
degeneration
and
necrosis
at
both
dose
levels.
The
incidence
of
renal
tubule
degeneration
was
greater
in
animals
receiving
the
aqueous
gavage
at
the
low
dose;
however,
the
severity
of
renal
degeneration
and
necrosis
was
greater
in
the
animals
receiving
the
oil
gavage
at
the
high
dose.
The
authors
attributed
the
vehicle
differences
to
slower
gastrointestinal
uptake
of
bromodichloromethane
from
the
oil
vehicle
compared
to
the
aqueous
vehicle.
At
the
high
dose,
more
bromodichloromethane
would
be
converted
to
a
reactive
metabolite
following
oil
dosing,
while
saturation
would
occur
following
aqueous
dosing.
At
the
low
dose,
the
difference
in
uptake
would
have
less
of
an
effect.
Overall,
this
study
found
that
the
kidney
was
more
sensitive
than
the
liver
to
a
single
dose
of
bromodichloromethane.
A
LOAEL
of
200
mg/
kg
was
identified
for
each
vehicle
based
on
minimal
renal
tubule
degeneration
and
changes
in
markers
of
renal
function.

Lilly
et
al.
(
1996)
investigated
the
effect
of
subchronic
pretreatment
with
corn
oil
on
the
toxicity
of
bromodichloromethane.
Prior
to
initiation
of
dosing
with
bromodichloromethane,
male
Draft
­
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or
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February
20,
2003
V
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5
Fischer
344
rats
(
6
animals/
group)
were
gavaged
with
oral
doses
of
corn
oil
or
water
for
six
weeks
(
5
days/
week)
at
a
constant
volume
of
5
mL/
kg.
Following
pretreatment,
the
animals
were
gavaged
with
a
single
dose
of
0,
200,
or
400
mg
bromodichloromethane/
kg
in
10%
Emulphor
®
.
Urine
was
collected
at
24,
36,
and
48
hours
following
bromodichloromethane
administration.
The
rats
were
sacrificed
at
48
hours
and
necropsies
were
performed.
Activities
of
the
hepatotoxicity
indicators
alanine
aminotransferase
(
ALT),
aspartate
aminotransferase
(
AST),
lactate
dehydrogenase
(
LDH),
and
sorbitol
dehydrogenase
(
SDH)
were
measured
in
the
serum,
and
the
renal
toxicity
indicators
alkaline
phosphatase
(
ALK),
AST,
and
LDH,
were
measured
in
the
urine.
Additional
analyses
included
determination
of
serum
levels
of
bile
acids,
triglycerides,
cholesterol
and
albumin,
and
urine
levels
of
N­
acetylglucosaminidase
and
gamma
glutamyl
transpeptidase
activity.
Enzymatic
activity
of
cytochrome
P450
isoforms
CYP2E1
and
CYP2B1/
B2
was
measured
in
the
microsomal
fraction
of
the
liver
to
investigate
whether
corn
oil
was
an
inducer
of
bromodichloromethane
metabolizing
enzymes.

Liver
weight
was
significantly
reduced
only
in
the
water
pretreatment
group
at
the
high
dose.
Kidney
weight
was
reduced
in
both
pretreatment
groups
at
the
high
dose.
Activities
of
serum
AST
and
LDH
were
significantly
elevated
in
both
pretreatment
groups
at
400
mg/
kg.
ALT
levels
increased
in
a
dose­
dependent
manner
in
the
water
pretreatment
group,
but
significant
elevations
were
noted
only
at
the
400
mg/
kg
dose
in
animals
pretreated
with
corn
oil.
Activities
of
urinary
AST
and
LDH
were
greater
than
controls
in
both
pretreatment
groups
after
24,
36,
and
48
hours.
ALK
levels
were
significantly
increased
in
both
pretreatment
groups
at
24
hours.
At
36
and
48
hours,
ALK
levels
were
elevated
only
in
water­
pretreated
animals.
High
incidences
of
renal
tubular
necrosis
occurred
at
200
and
400
mg/
kg
in
both
pretreatment
groups.
There
were
no
significant
differences
in
the
histopathological
lesion
scores
between
the
pretreatment
groups.
No
significant
differences
were
noted
in
the
hepatic
activity
of
CYP2E1
or
CYP2B1/
B2
in
the
corn
oil
pretreated
animals
compared
to
the
water
controls.
Although
a
number
of
differences
between
the
pretreatment
groups
were
noted
in
results
for
specific
endpoints,
the
overall
results
from
this
study
indicate
that
6
weeks
of
pretreatment
with
corn
oil
did
not
significantly
enhance
the
acute
hepato­
or
nephrotoxicity
of
bromodichloromethane.
In
addition,
the
reported
data
suggest
that
vehicle­
related
differences
in
toxicity
observed
in
other
bromodichloromethane
studies
are
most
likely
due
to
pharmacokinetic
differences
in
absorption
rather
than
altered
enzyme
activity
induced
by
corn
oil.
This
study
confirms
the
acute
LOAEL
of
200
mg/
kg­
day
previously
identified
by
Lilly
et
al.
(
1994)
for
renal
toxicity.

Lilly
et
al.
(
1997)
administered
single
doses
of
bromodichloromethane
by
gavage
in
aqueous
10%
Emulphor
®
solution
to
male
F344
rats
at
dose
levels
of
0,
123,
164,
246,
328,
or
492
mg/
kg.
Groups
of
5
animals/
dose
were
sacrificed
at
24
and
48
hours
post­
dosing.
Body
weights
were
significantly
decreased
at
or
above
246
mg/
kg
after
48
hours.
At
24
hours,
absolute
and
relative
kidney
weights
were
significantly
increased
at
or
above
328
mg/
kg
and
246
mg/
kg,
respectively.
At
48
hours,
only
relative
kidney
weight
at
the
high
dose
was
significantly
increased.
At
24
hours,
serum
markers
of
liver
damage
(
ALT
and
AST)
were
significantly
increased
at
or
above
246
mg/
kg
with
one
marker
(
SDH)
increased
at
all
dose
levels.
Although
smaller
statistically
significant
increases
were
observed
at
the
low
doses
at
24
hours
for
ALT
(
123
and
164
mg/
kg)
and
AST
(
164
mg/
kg),
the
biological
significance
of
these
increases
is
unclear.
After
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February
20,
2003
V
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6
48
hours,
serum
levels
of
these
markers
were
decreased
from
24­
hour
levels
with
statistically
significant
changes
noted
only
at
the
higher
doses.
No
effects
in
urinary
markers
of
kidney
damage
were
found
at
either
123
or
164
mg/
kg.
These
markers,
however,
were
significantly
elevated
after
24
hours
for
doses
at
or
above
246
mg/
kg
with
few
exceptions.
No
histopathological
examination
was
conducted.
These
results
were
generally
consistent
with
earlier
results
(
Lilly
et
al.,
1994),
although
the
present
study
was
conducted
at
doses
low
enough
to
identify
a
NOAEL.
In
contrast
to
the
earlier
results
of
Lilly
et
al.
(
1994),
this
study
did
not
find
that
the
kidney
was
more
sensitive
than
the
liver
to
the
toxic
effects
of
bromodichloromethane.
Based
on
hepatotoxicity
and
nephrotoxicity,
this
study
identified
a
NOAEL
of
164
mg/
kg
and
a
LOAEL
of
246
mg/
kg.

Keegan
et
al.
(
1998)
investigated
the
acute
toxicity
of
bromodichloromethane
administered
orally
in
an
aqueous
vehicle.
Male
Fischer
344
rats
(
6
animals/
group)
were
gavaged
with
a
single
dose
of
0,
0.125,
0.1875,
0.250,
0.5,
0.75,
1.0
or
1.5
mmol/
kg
dissolved
in
a
10%
aqueous
solution
of
Alkamuls
EL­
620.
These
doses
of
bromodichloromethane
are
equivalent
to
0,
20.5,
30.7,
41.0,
81.9,
122.9,
163.8,
and
245.7
mg/
kg,
respectively.
Control
animals
were
dosed
with
vehicle
only
(
10%
Alkamuls
EL­
620).
Gavage
volumes
were
kept
constant
at
5
ml/
kg
body
weight.
Animals
were
sacrificed
24
hours
after
dose
administration
and
the
liver,
kidneys,
and
serum
were
harvested.
Significant
decreases
in
body
weight
were
observed
in
the
0.75,
1.0,
and
1.5
mmol
bromodichloromethane/
kg
treated
animals.
Decreases
in
absolute
liver
weights
were
observed
in
the
0.5,
0.75,
1.0,
and
1.5
mmol
bromodichloromethane/
kg
treated
animals.
No
change
was
noted
in
relative
liver
weights.
Absolute
kidney
weights
were
not
effected
by
bromodichloromethane
treatment,
but
relative
kidney
weights
were
significantly
increased
in
the
two
highest
dose
groups
(
1.0
and
1.5
mmol
bromodichloromethane/
kg).
Serum
levels
of
ALT,
SDH,
and
AST
were
assessed
as
an
indication
of
liver
toxicity.
Dose­
dependent
elevations
in
ALT
(
45%
to
239%
increase),
AST
(
25%
to
130%
increase)
and
SDH
(
74%
to
378%
increase)
were
observed
in
the
0.5,
0.75,
1.0,
and
1.5
mmol
dose
groups.
Based
on
these
findings,
0.25
mmol/
kg
(
41.0
mg/
kg)
represents
the
NOAEL
and
0.5
mmol/
kg
(
81.9
mg/
kg)
represents
the
LOAEL
for
orally
administered
bromodichloromethane
in
an
aqueous
vehicle.
These
authors
used
the
NOAEL
of
41.0
mg/
kg
to
calculate
One­
Day
Health
Advisories
for
drinking
water
of
4
mg/
L
for
a
10­
kg
child
and
14
mg/
L
for
a
70­
kg
adult.

2.
Dibromochloromethane
The
acute
oral
lethality
of
dibromochloromethane
has
been
assessed
in
rats
and
mice
of
both
sexes.
Chu
et
al.
(
1980)
reported
LD
50
values
in
male
and
female
Sprague­
Dawley
rats
of
1,186
and
848
mg/
kg­
day
for
males
and
females,
respectively.
Bowman
et
al.
(
1978)
reported
LD
50
values
in
mice
of
800
and
1,200
mg/
kg­
day,
respectively.

Bowman
et
al.
(
1978)
investigated
the
acute
oral
toxicity
of
dibromochloromethane
in
ICR
Swiss
mice
(
10/
sex/
group).
Doses
of
500
to
4000
mg/
kg
(
individual
doses
not
reported)
were
administered
by
gavage
in
Emulphor
®
:
alcohol:
saline
(
1:
1:
8)
to
fasted
animals.
Sedation
and
anesthesia
occurred
at
500
mg/
kg.
Males
were
more
sensitive
than
females
to
the
acute
lethal
effects
of
dibromochloromethane.
Draft
­
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February
20,
2003
V
­
7
NTP
(
1985)
evaluated
the
acute
toxicity
of
dibromochloromethane
in
male
and
female
F344/
N
rats.
The
rats
(
5
animals/
sex/
dose)
received
single
doses
of
160,
310,
630,
1,250,
or
2,500
mg/
kg
dibromochloromethane
by
gavage
in
corn
oil.
The
observation
period
following
treatment
was
14
days.
Mortality
in
high­
dose
rats
was
100%
by
day
3.
At
the
1,250
mg/
kg
dose,
four
male
rats
and
one
female
rat
died.
One
female
rat
died
in
the
630
mg/
kg
group.
Doses
of
310
mg/
kg
or
greater
produced
lethargy
in
all
animals
for
3
hours
after
dosing.
A
gross
necropsy
was
conducted
on
one
or
two
animals
from
each
group.
No
treatment­
related
effects
were
observed
in
rats
selected
for
gross
necropsy.

In
a
concurrent
study,
NTP
(
1985)
evaluated
the
acute
toxicity
of
dibromochloromethane
in
male
and
female
B6C3F
1
mice
(
5/
sex/
dose).
The
mice
received
single
doses
of
160,
310,
630,
1,250,
or
2,500
mg/
kg
dibromochloromethane
by
gavage
in
corn
oil.
The
observation
period
following
treatment
was
14
days.
All
male
mice
receiving
the
2,500
mg/
kg
and
1,250
mg/
kg
doses
died.
Three
male
mice
receiving
the
630
mg/
kg
dose
died,
while
a
single
male
mouse
died
at
the
310
mg/
kg
dose.
All
female
mice
receiving
the
2500
mg/
kg
dose
died.
Four
of
the
female
mice
administered
the
1,250
mg/
kg
dose
died
between
days
2
and
8
post­
treatment.
No
female
mice
died
at
doses
of
630
mg/
kg
or
lower.
A
gross
necropsy
was
conducted
on
one
or
two
animals
from
each
group.
At
necropsy,
aberrations
of
the
kidney
(
dark
red
or
pale
medullae)
and
liver
(
discolored
foci)
were
reported
to
be
more
frequently
observed
in
treated
animals
than
in
control
animals
(
raw
data
were
not
presented
in
the
study).

Müller
et
al.
(
1997)
investigated
the
cardiotoxic
effects
of
acute
dibromochloromethane
exposure.
Male
Wistar
rats
were
administered
a
single
dose
of
dibromochloromethane
by
gavage
in
olive
oil
at
dose
levels
of
0,
83,
167,
333,
or
667
mg/
kg.
Telemetric
measurements
of
cardiovascular
parameters
(
heart
rate,
blood
pressure,
body
temperature,
and
physical
activity)
were
recorded
in
conscious
rats
(
6/
group)
24
hours
prior
to
administration
to
72
hours
following
administration.
Heart
rate
and
blood
pressure
were
also
measured
in
urethane­
anesthetized
rats
(
10/
group)
25
minutes
following
administration.
For
these
rats,
contractility
parameters,
such
as
the
Krayenbühl
index,
were
also
calculated.
Treatment­
related
arrhythmias
were
not
observed
in
conscious
rats
dosed
with
83
to
333
mg/
kg
of
dibromochloromethane
while
rats
in
the
high­
dose
group
exhibited
premature
ventricular
contractions
one
minute
following
administration.
Heart
rate
and
body
temperature
were
initially
decreased
in
all
treatment
groups
following
administration,
but
returned
to
control
values
24
hours
post­
exposure
in
rats
administered
83
to
333
mg/
kg.
In
the
high­
dose
rats,
heart
rate
remained
depressed
up
to
48
hours
post­
exposure,
and
body
temperature
decreased
4.5

C
below
control
values
by
72
hours
post­
exposure.
Blood
pressure
was
initially
increased
in
all
treatment
groups
following
administration,
but
began
to
return
to
control
values
within
48
hours
post­
exposure
in
rats
administered
83
to
333
mg/
kg.
Blood
pressure
in
the
high­
dose
group,
however,
decreased
below
control
values
72
hours
postexposure
Physical
activity
was
decreased
in
conscious
rats
administered
333
and
667
mg/
kg
during
the
entire
observation
period.
In
urethane­
anesthetized
rats,
negative
effects
on
muscle
contractility
were
observed
at
dose
levels
of
333
and
667
mg/
kg,
negative
chronotropic
(
rate
of
contraction)
effects
were
observed
at
the
333
mg/
kg
dose
level,
and
negative
dromotropic
(
defined
as
influencing
the
velocity
of
conduction
of
excitation,
as
in
nerve
or
cardiac
muscle
Draft
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February
20,
2003
V
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8
fibers)
effects
were
observed
at
dose
levels
of
167
to
667
mg/
kg.
Heart
rate,
blood
pressure,
and
several
contractility
parameters,
however,
did
not
exhibit
dose­
related
trends.

3.
Bromoform
Bowman
et
al.
(
1978)
assessed
the
acute
oral
toxicity
of
bromoform
in
ICR
Swiss
mice.
Groups
of
ten
male
(
30
to
35
g)
and
ten
female
(
25
to
30
g)
mice
were
treated
with
single
doses
ranging
from
500
to
4,000
mg/
kg.
Compounds
were
solubilized
in
Emulphor
®
:
alcohol:
saline
(
1:
1:
8)
and
administered
by
gavage
to
fasted
animals.
The
period
of
observation
following
treatment
was
14
days.
LD
50
values
were
1400
and
1550
mg/
kg
for
males
and
females,
respectively.
Ataxia,
sedation,
and
anesthesia
occurred
within
60
minutes
of
treatment
at
doses
of
1000­
mg/
kg
and
above.
Sedation
lasted
approximately
4
hours.

Chu
et
al.
(
1980)
evaluated
the
acute
toxicity
of
bromoform
in
male
and
female
Sprague­
Dawley
rats.
Fasted
adult
rats
(
10/
sex/
dose)
received
doses
of
546,
765,
1071,
1500,
or
2100
mg/
kg
bromoform
dissolved
in
corn
oil
by
gavage.
Clinical
observations
were
made
for
14
days
after
treatment.
The
LD
50
values
for
male
and
female
rats
were
1388
and
1147
mg/
kg,
respectively.
Clinical
signs
observed
in
treated
rats
included
sedation,
flaccid
muscle
tone,
ataxia,
piloerection,
and
hypothermia.
Gross
pathological
examination
revealed
liver
and
kidney
congestion
in
treated
animals.
Chu
et
al.
(
1982a)
reported
results
for
growth,
food
intake,
organ
weight,
histopathology,
hematological
indices,
liver
microsome
aniline
hydroxylase
activity
and
serum
chemistry
in
surviving
rats.
Bromoform
treatment
increased
liver
protein
concentration
in
the
serum
of
male
rats
at
doses
of
765
and
1071
mg/
kg.
Lymphocyte
counts
were
decreased
in
male
(
765
and
1071
mg/
kg
doses)
and
female
(
765
mg/
kg)
rats
but
the
effect
was
not
dosedependent
Female
rats
at
the
765
mg/
kg
dose
had
elevated
aniline
hydroxylase
levels.

NTP
(
1989a)
investigated
the
acute
oral
toxicity
of
bromoform
in
male
and
female
F344/
N
rats.
The
rats
(
5/
sex/
group)
were
administered
a
single
oral
dose
of
bromoform
(
by
gavage,
in
corn
oil)
at
dose
levels
of
125,
250,
500,
1,000,
or
2,000
mg/
kg.
Control
groups
were
not
included
in
the
study
design.
Mortality
was
10/
10
at
2,000
mg/
kg,
6/
10
at
1,000
mg/
kg,
and
0/
10
at
500
mg/
kg
or
lower.
Shallow
breathing
was
observed
in
rats
that
received
the
1000
or
2000
mg/
kg
doses.
No
other
clinical
signs
were
reported.

NTP
(
1989a)
investigated
the
acute
oral
toxicity
of
bromoform
in
male
and
female
B6C3F
1
mice.
The
mice
received
a
single
oral
dose
of
bromoform
(
by
gavage,
in
corn
oil)
at
dose
levels
of
125,
250,
500,
1,000,
or
2,000
mg/
kg.
There
were
no
controls.
Mortality
was
0/
10
at
2,000
mg/
kg,
6/
10
at
1,000
mg/
kg,
1/
10
at
500
mg/
kg,
and
0/
10
at
250
mg/
kg
or
lower.
The
final
mean
body
weight
of
mice
that
survived
to
the
end
of
the
study
period
was
unaffected
by
bromoform
exposure.
Male
mice
that
received
doses
of
500,
1,000,
or
2,000
mg/
kg
and
females
that
received
1,000
or
2,000
mg/
kg
were
lethargic.
Shallow
breathing
was
noted
in
male
mice
administered
the
1,000
or
2,000
mg/
kg
dose.

B.
Short­
Term
Exposures
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
9
This
section
summarizes
short­
term
studies
(
less
than
approximately
90
days)
on
the
health
effects
of
brominated
trihalomethanes
in
animals.
Details
of
these
studies
are
summarized
in
Table
V­
3.

Table
V­
3
Summary
of
Short
Term
Toxicity
Studies
for
Brominated
Trihalomethanes
Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Bromodichloromethane
Oral
Exposure
Chu
et
al.
(
1982a)
Rat
SD*
Drinking
water
M
10
28
days
0
0.8
8
68
(
NOAEL)
No
signs
of
toxicity
observed.

Munson
et
al.
(
1982)
Mouse
CD­
1
Gavage
(
aqueous)
M,
F
8­
12
14
days
0
50
(
NOAEL)
125
(
LOAEL)
250
Decreased
immune
function;
increased
liver
weight,
decreased
absolute
and
relative
spleen
wt.
(
females)

Condie
et
al.
(
1983)
Mouse
CD­
1
Gavage
(
corn
oil)
M
8­
16
14
days
0
37
74
(
NOAEL)
148
(
LOAEL)
Liver
and
kidney
histopathology
NTP
(
1987)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
5
14
days
0
38
75
150
(
NOAEL)
300
(
LOAEL)
600
Decreased
body
weight
gain;
renal
pathology
NTP
(
1987)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
5
14
days
0
19
38
75
(
NOAEL)
150
(
LOAEL)
300
Mortality,
renal
histopathology
Aida
et
al.
(
1992a)
Rat
Wistar
Diet
M
7
1
month
0
21
62
(
NOAEL)
189
(
LOAEL)
Liver
histopathology
Aida
et
al.
(
1992a)
Rat
Wistar
Diet
F
7
1
month
0
21
66
(
NOAEL)
204
(
LOAEL)
Liver
histopathology
Table
V­
3
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
10
Thornton­
Manning
et
al.
(
1994)
Rat
F344
Gavage
(
aqueous)
F
6
5
days
0
75
(
NOAEL)
150
(
LOAEL)
300
Liver
histopathology,
renal
histopathology;
increased
liver
and
kidney
wt.;
elevated
markers
of
hepatotoxicity
Thornton­
Manning
et
al.
(
1994)
Mouse
C57BL/
6J
Gavage
(
aqueous)
F
6
5
days
0
75
(
NOAEL)
150
(
LOAEL)
Increased
serum
markers
of
hepatotoxicity
Potter
et
al.
(
1996)
Rat
F344
Gavage
(
aqueous)
M
1,
3,
or
7
days
123
246
(
NOAEL)
No
effect
in
hyaline
droplet
formation
or
cell
proliferation
Melnick
et
al.
(
1998)
Mouse
B6C3F1
Gavage
(
corn
oil)
F
10
3
weeks
(
5
d/
wk)
0
75
(
NOAEL)
150
(
LOAEL)
326
Increased
abs.
and
relative
liver
weight;
increased
serum
markers
of
hepatotoxicity;
hepatocyte
degeneration;
increased
labeling
index
NTP
(
1998)
Rat
SD
Drinking
water
M,
F
6
2
weeks
0
11
45
(
NOAEL)
91
(
LOAEL)
124
Transient
reduction
in
weight
gain
NTP
(
1998)
Rat
SD
Drinking
water
M,
F
5­
13
35
days
Group
A
males
0
9
(
NOAEL)
38
(
LOAEL)
67
Single
cell
hepatic
necrosis
in
Group
A
males
Coffin
et
al.
(
2000)
Mouse
B6C3F1
Gavage
(
Corn
oil)

Drinking
water
F
10
11
days
11
days
0
150
(
LOAEL)
300
0
138
(
LOAEL)
Hydropic
degeneration
in
liver
(
corn
oil
gavage
and
drinking
water);
increased
relative
liver
weight
(
gavage);
increased
proliferating
cell
nuclear
antigen
labeling
index
(
gavage)

Inhalation
Exposure
Torti
et
al.
(
2001)
Mouse
C57BL/
6
FVB/
N
(
wildtype
Vapor
M
6
1
week
(
6
hr/
day)
0
ppm
1
ppm
(
NOAEL)
10
ppm
(
LOAEL)
30
ppm
100
ppm
150
ppm
Dose­
dependent
marginal
increase
in
renal
tubular
degeneration
in
C57BL/
6
mice;
mild
increase
in
renal
tubular
degeneration
and
marginal
increase
in
hepatic
degeneration
in
FVB/
N
mice;
sign.
increased
labeling
index
Table
V­
3
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
11
Torti
et
al.
(
2001)
Mouse
C57BL/
6
(
p53
heterozygous
Vapor
M
6
1
week
(
6
hr/
day)
0
1
ppm
(
NOAEL)
10
ppm
(
LOAEL)
30
ppm
100
ppm
150
ppm
Dose­
dependent
marginal
to
mild
increase
in
renal
tubular
degeneration;
sign.
increased
labeling
index
Torti
et
al.
(
2001)
FVB/
N
(
p53
heterozygous
Vapor
M
6
1
week
(
6
hr/
day)
0
ppm
0.3
ppm
1
ppm
3
ppm
(
NOAEL)
10
ppm
(
LOAEL)
30
ppm
Dose­
dependent
mild
increase
in
renal
tubular
degeneration
and
marginal
increase
in
nephrosis;
marginal
increase
in
hepatic
degeneration;
sign.
increased
relative
kidney
wt.
and
labeling
index.

Torti
et
al.
(
2001)
Mouse
C57BL/
6
FVB/
N
(
wildtype
Vapor
M
6
3
weeks
(
6
hr/
day)
0
ppm
0.3
ppm
1
ppm
3
ppm
(
NOAEL)
10
ppm
(
LOAEL)
30
ppm
Marginal
increase
in
renal
tubular
degeneration
Torti
et
al.
(
2001)
Mouse
C57BL/
6
FVB/
N
(
p53
heterozygous
Vapor
M
6
3
weeks
(
6
hr/
day)
0
ppm
0.3
ppm
1
ppm
3
ppm
(
NOAEL)
10
ppm
(
LOAEL)
30
ppm
Marginal
or
mild
increase
in
renal
tubular
degeneration
in
both
strains;
marginal
increase
in
hepatic
degeneration
in
FVB/
N
heterozygous
strain
Dibromochloromethane
Munson
et
al.
(
1982)
Mouse
CD­
1
Gavage
(
aqueous)
M,
F
8­
12
14
days
0
50
(
NOAEL)
125
(
LOAEL)
250
Decreased
immune
function
Chu
et
al.
(
1982a)
Rat
SD
Drinking
water
M
10
28
days
0
0.7
8.5
68
(
NOAEL)
No
effect
on
growth,
clinical
signs,
biochemical
or
histopathological
endpoints
Condie
et
al.
(
1983)
Mouse
CD­
1
Gavage
(
corn
oil)
M
8­
16
14
days
0
37
74
(
NOAEL)
147
(
LOAEL)
Decreased
PAH
uptake,
moderate
liver
and
kidney
histopathology
Table
V­
3
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
12
NTP
(
1985)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
5
14
days
0
60
125
250
(
NOAEL)
500
(
LOAEL)
1,000
Mortality;
liver
and
renal
gross
pathology
NTP
(
1985)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
5
14
days
0
30
60
(
NOAEL)
125
(
LOAEL)
250
500
Liver
and
kidney
gross
pathology
Aida
et
al.
(
1992a)
Rat
Wistar
Diet
M
7
1
month
0
18
(
NOAEL)
56
(
LOAEL)
173
Liver
histopathology
Aida
et
al.
(
1992a)
Rat
Wistar
Diet
F
7
1
month
0
34
(
NOAEL)
101
(
LOAEL)
333
Liver
histopathology;
increased
relative
liver
weight
Potter
et
al.
(
1996)
Rat
F344
Gavage
(
aqueous)
M
1,
3,
or
7
days
156
312
(
NOAEL)
No
effect
on
hyaline
droplet
formation
or
cell
proliferation
Melnick
et
al.
(
1998)
Mouse
B6C3F1
Gavage
(
corn
oil)
F
10
3
weeks
(
5
d/
wk)
0
50
100
(
NOAEL)
192
(
LOAEL)
417
Liver
histopathology;
increased
serum
enzymes
and
liver
weight
Coffin
et
al.
(
2000)
Mouse
B6C3F1
Gavage
(
Corn
oil)

Drinking
water
F
10
11
days
11
days
0
100
(
LOAEL)
300
0
171
Increased
proliferating
cell
nuclear
antigen
labeling
index
(
gavage);
increased
relative
liver
wt.

Bromoform
Munson
et
al.
(
1982)
Mouse
CD­
1
Gavage
(
aqueous)
M,
F
6­
12
14
days
0
50
125
(
NOAEL)
250
(
LOAEL)
Increased
serum
enzyme
activity
(
AST);
decrease
in
antibody
forming
cells
and
delayed­
type
hypersensitivity
response
Chu
et
al.
(
1982a)
Rat
SD
Drinking
water
M
10
28
days
0.7
8.5
80
(
NOAEL)
No
effect
on
growth,
clinical
signs,
biochemical
or
histopathological
endpoints
Table
V­
3
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
13
Condie
et
al.
(
1983)
Mouse
CD­
1
Gavage
(
corn
oil)
M
8­
16
14
days
0
72
145
(
NOAEL)
289
(
LOAEL)
Decreased
PAH
uptake,
moderate
histopathological
changes
NTP
(
1989a)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
5
14
days
0
100
200
(
NOAEL)
400
(
LOAEL)
600
800
Decreased
body
weight
gain;
1/
5
died
at
400
mg/
kgday
100%
mortality
at
two
highest
doses
NTP
(
1989a)
Mouse
B6C3F1
Gavage
(
corn
oil)
M
5
14
days
0
50
100
200
(
NOAEL)
400
(
LOAEL)
600
Stomach
nodules;
ataxia,
lethargy;
1/
5
died
at
high
dose
Aida
et
al.
(
1992a)
Rat
Wistar
Diet
M
7
1
month
0
62
(
NOAEL)
187
(
LOAEL)
618
Hepatic
vacuolization,
serum
chemistry/
biochemistry
Aida
et
al.
(
1992a)
Rat
Wistar
Diet
F
7
1
month
0
56
(
NOAEL)
208
(
LOAEL)
728
Hepatic
vacuolization,
serum
chemistry/
biochemistry
Potter
et
al.
(
1996)
Rat
F344
Gavage
(
aqueous)
M
1,
3,
or
7
days
190
379
(
NOAEL)
No
effect
on
hyaline
droplet
formation
or
cell
proliferation
Melnick
et
al.
(
1998)
Mouse
B6C3F1
Gavage
(
corn
oil)
F
10
3
weeks
(
5
d/
wk)
0
200
(
NOAEL)
500
(
LOAEL)
Increase
in
absolute
and
relative
liver
wt.;
marginally
significant
increase
in
LI
at
highest
dose
Coffin
et
al.
(
2000)
Mouse
B6C3F1
Gavage
(
corn
oil)

Drinking
Water
F
10
11
days
11
days
0
200
(
LOAEL)
500
0
301
Liver
histopathology;
increased
proliferating
cell
nuclear
antigen
labeling
index;
increased
relative
liver
wt.

*
SD,
Sprague­
Dawley
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
14
1.
Bromodichloromethane
Munson
et
al.
(
1982)
administered
bromodichloromethane
by
aqueous
gavage
to
male
and
female
CD­
1
mice
(
8
to
12/
sex/
group)
for
14
days
at
levels
of
0,
50,
125,
or
250
mg/
kg­
day.
Endpoints
evaluated
included
body
and
organ
weights,
hematology,
serum
enzyme
levels
(
SGOT,
SGPT),
and
humoral
and
cell­
mediated
immune
system
functions.
At
250
mg/
kg­
day,
body
weights
were
significantly
decreased.
Significant
organ
weight
changes
included
increased
relative
liver
weight
(
mid­
and
high­
dose
groups),
decreased
absolute
spleen
weight
(
high­
dose
males
and
mid­
and
high­
dose
females),
and
decreased
relative
spleen
weight
(
mid­
and
high­
dose
females).

Among
the
hematology
endpoints,
only
fibrinogen
levels
were
significantly
decreased
in
high­
dose
males
and
in
mid­
and
high­
dose
females.
Significant
clinical
chemistry
findings
included
decreased
glucose
levels
(
high­
dose
males),
increased
ALT
and
AST
activities
(
highdose
groups),
and
increased
blood
urea
nitrogen
(
BUN)
levels
(
high­
dose
groups).
Bromodichloromethane
appeared
to
affect
the
humoral
immune
system,
as
judged
by
significantly
decreased
antibody­
forming
cells
(
high­
dose
males
and
mid­
and
high­
dose
females)
and
hemagglutination
titers
(
mid­
and
high­
dose
males
and
high­
dose
females).
This
study
identified
a
NOAEL
of
50
mg/
kg­
day
and
a
LOAEL
of
125
mg/
kg­
day
for
bromodichloromethane
on
the
basis
of
decreased
immune
function
in
females.

Chu
et
al.
(
1982a)
administered
bromodichloromethane
to
male
Sprague­
Dawley
rats
(
10/
group)
in
drinking
water
for
28
days
at
dose
levels
of
0,
5,
50,
or
500
ppm.
These
levels
corresponded
to
doses
of
0,
0.8,
8.0,
or
68
mg/
kg­
day,
as
calculated
by
the
authors
based
on
recorded
fluid
intake.
The
authors
observed
no
effects
on
growth
rate
or
food
consumption
and
no
signs
of
toxicity
throughout
the
exposure.
No
dose­
related
biochemical
or
histologic
changes
were
detected
(
no
data
were
provided).
This
study
identified
a
NOAEL
of
68
mg/
kg­
day,
but
the
reported
data
were
too
limited
to
allow
an
independent
verification.

Condie
et
al.
(
1983)
investigated
the
renal
and
hepatic
toxicity
of
bromodichloromethane
in
male
CD­
1
mice
(
8
to
16/
group).
Bromodichloromethane
was
administered
by
gavage
in
corn
oil
for
14
days
at
dose
levels
of
0,
37,
74
or
148
mg/
kg­
day.
Biochemical
evidence
of
liver
damage
(
significantly
elevated
ALT)
was
observed
at
the
high
dose,
while
biochemical
evidence
of
kidney
damage
(
significantly
decreased
p­
aminohippurate
(
PAH)
uptake
by
kidney
slices)
was
observed
at
the
mid
and
high
dose.
Significantly
decreased
BUN
levels
were
observed
in
the
lowand
mid­
dose
groups,
but
not
in
the
high­
dose
group.
Histopathology
revealed
no
consistent
or
important
changes
at
the
low
or
mid­
level
doses,
with
minimal
to
moderate
liver
and
kidney
injury
observed
in
the
majority
of
animals
at
the
high
dose.
Liver
lesions
included
centrilobular
pallor
and
focal
inflammation.
Kidney
lesions
included
intratubular
mineralization,
epithelial
hyperplasia,
and
cytomegaly.
Although
the
severity
of
these
lesions
was
primarily
minimal
to
slight,
a
few
animals
in
the
high
dose
group
exhibited
moderate
to
moderately
severe
intratubular
mineralization
and/
or
epithelial
hyperplasia.
This
study
identified
a
NOAEL
value
of
74
mg/
kgday
and
a
LOAEL
value
of
148
mg/
kg­
day
for
bromodichloromethane,
based
on
histopathology.
Draft
­
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February
20,
2003
V
­
15
NTP
(
1987)
administered
doses
of
0,
38,
75,
150,
300,
or
600
mg/
kg­
day
of
bromodichloromethane
in
corn
oil
by
gavage
to
male
and
female
F344/
N
rats
(
5/
sex/
dose)
for
14
days.
One
low­
dose
and
one
high­
dose
female
died
before
study
termination.
All
high­
dose
animals
were
hyperactive
after
dosing
and
either
lost
weight
or
gained
no
weight
during
the
study.
Final
mean
body
weights
were
not
significantly
affected
in
groups
given
38,
75,
or
150
mg/
kgday
At
300
mg/
kg,
body
weights
of
males
and
females
were
decreased
by
21%
and
7%,
respectively,
relative
to
vehicle
controls.
At
600
mg/
kg­
day,
body
weights
of
males
and
females
were
decreased
by
44%
and
22%,
respectively,
relative
to
vehicle
controls.
Necropsy
was
performed
on
all
animals.
Renal
medullae
were
reddened
in
all
high­
dose
males
and
in
one
female
in
each
of
the
control,
low­
dose,
and
high­
dose
groups.
This
study
identified
a
NOAEL
of
150
mg/
kg­
day
and
a
LOAEL
of
300
mg/
kg­
day
in
rats,
based
on
decreased
body
weight
gain.
In
a
parallel
experiment,
NTP
(
1987)
administered
doses
of
0,
19,
38,
75,
150,
or
300
mg/
kg­
day
bromodichloromethane
in
corn
oil
by
gavage
to
male
and
female
B6C3F
1
mice
(
5/
sex/
dose)
for
14
days.
All
male
mice
that
received
150
or
300
mg/
kg­
day
bromodichloromethane
died
before
study
termination.
Clinical
signs
included
lethargy,
dehydration,
and
hunched
posture.
The
final
mean
body
weights
of
the
mice
that
survived
were
not
significantly
different
from
the
controls.
The
renal
medullae
were
reddened
in
four
males
in
the
150
mg/
kg­
day
group,
all
males
in
the
300
mg/
kg­
day
group,
and
one
female
in
the
150
mg/
kg­
day
group.
Based
on
behavior,
appearance,
gross
necropsy,
and
mortality,
this
study
identified
a
NOAEL
of
75
mg/
kg­
day
and
a
frank
effect
level
(
FEL)
of
150
mg/
kg­
day
in
male
mice.
An
interesting
point
to
note
is
that
this
study
and
the
study
by
Condie
et
al.
(
1983)
were
conducted
under
similar
conditions
(
mice
administered
bromodichloromethane
by
gavage
in
corn
oil
for
14
days),
but
with
dramatically
different
results.
In
contrast
to
the
100%
mortality
observed
in
this
study
for
male
mice,
Condie
et
al.
(
1983)
found
only
moderate
histopathology
in
male
CD­
1mice
at
148
mg/
kg­
day
with
no
deaths
occurring.
The
reason
for
this
difference
is
unclear,
but
may
be
related
to
strain­
specific
differences
in
sensitivity.

Aida
et
al.
(
1992a)
administered
bromodichloromethane
to
Slc:
Wistar
rats
(
7/
sex/
group)
for
one
month
at
dietary
levels
of
0%,
0.024%,
0.072%,
or
0.215%
for
males
and
0%,
0.024%,
0.076%,
or
0.227%
for
females.
The
test
material
was
microencapsulated
and
mixed
with
powdered
feed;
placebo
granules
were
used
for
the
control
groups.
Based
on
the
mean
food
intakes,
the
study
authors
reported
the
mean
compound
intakes
for
the
one­
month
period
as
0,
20.6,
61.7,
or
189.0
mg/
kg­
day
for
males
and
0,
21.1,
65.8,
or
203.8
mg/
kg­
day
for
females.
Clinical
effects,
body
weight,
food
consumption,
hematology
parameters,
serum
chemistry,
and
histopathology
of
all
major
organs
were
determined.
Body
weights
were
significantly
decreased
in
the
high­
dose
groups
relative
to
the
controls.
The
high­
dose
animals
also
exhibited
slight
piloerection
and
emaciation.
Relative
liver
weight
was
increased
only
in
high­
dose
females.
Significant,
dose­
related
biochemical
findings
at
the
low
dose
were
limited
to
decreased
LDH
levels
in
males,
but
the
biological
significance
of
this
effect
is
unclear.
Serum
LDH
levels
were
also
significantly
decreased
at
the
low
and
high
dose
in
females.
Other
statistically
significant,
dose­
related
changes
included
decreased
glucose
(
high­
dose
males),
decreased
serum
triglycerides
(
high­
dose
groups),
decreased
serum
cholinesterase
activity
(
high­
dose
males
and
mid­
and
highdose
females),
and
increased
total
cholesterol
(
mid­
and
high­
dose
males).
The
changes
in
cholinesterase
activity
and
cholesterol
levels
in
males
were
not
dose­
related.
The
cholesterol
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February
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V
­
16
levels
were
within
normal
ranges
at
all
doses.
Treatment­
related
histopathological
lesions
were
limited
to
the
liver
and
were
rated
as
very
slight
or
slight.
The
lesions
were
mostly
confined
to
the
high­
dose
groups.
Vacuolization
observed
in
mid­
dose
females
and
in
a
single
low­
dose
male
was
not
considered
an
adverse
effect.
Other
observed
effects
included
swelling
of
hepatocytes,
single
cell
necrosis,
hepatic
cord
irregularity,
and
bile
duct
proliferation.
These
lesions
were
observed
only
in
high­
dose
males
and
females
with
the
exception
of
very
slight
to
slight
changes
in
individual
low­
dose
males.
No
effect
was
observed
on
any
hematology
parameter.
Based
on
the
histopathology
observed
in
high­
dose
males
and
females,
the
LOAELs
identified
in
this
study
for
bromodichloromethane
in
rats
were
189.0
mg/
kg­
day
in
males
and
203.8
mg/
kg­
day
in
females;
the
NOAELs
were
61.7
mg/
kg­
day
in
males
and
65.8
mg/
kg­
day
in
females.

Thornton­
Manning
et
al.
(
1994)
administered
bromodichloromethane
at
dose
levels
of
0,
75,
150,
or
300
mg/
kg­
day
by
gavage
to
female
F344
rats
(
6
animals/
dose)
for
five
consecutive
days.
The
dosing
vehicle
consisted
of
an
aqueous
10%
Emulphor
®
solution.
Animals
were
sacrificed
on
day
6.
Two
animals
in
the
high­
dose
group
died
on
day
5.
Final
body
weights
of
the
high­
dose
group
were
significantly
decreased
compared
to
the
controls.
Absolute
and
relative
kidney
and
liver
weights
were
significantly
increased
at
150
and
300
mg/
kg­
day
with
the
exception
of
a
nonsignificant
increase
in
absolute
liver
weight
at
150
mg/
kg­
day.
Toxic
effects
on
the
kidney
and
liver
were
reflected
in
significantly
increased
LDH,
AST,
SDH,
creatinine,
and
BUN
at
300
mg/
kg­
day.
These
results
were
supported
by
the
histopathology
findings.
In
the
liver,
centrilobular
vacuolar
degeneration
was
observed
at
both
150
and
300
mg/
kg­
day
with
the
severity
of
the
effect
increased
with
increasing
dose.
Centrilobular
hepatocellular
necrosis
was
also
observed
in
one
high­
dose
animal.
In
the
kidney,
renal
tubular
vacuolar
degeneration
and
renal
tubule
regeneration
were
observed
at
150
and
300
mg/
kg­
day
with
the
incidence
and
severity
increased
with
increasing
dose.
While
minimal
renal
tubule
necrosis
was
observed
in
only
one
animal
at
the
mid
dose,
all
animals
at
the
high
dose
exhibited
mild
to
moderate
renal
tubule
necrosis.
Significant
decreases
in
the
hepatic
activity
of
the
CYP1A
and
CYP2B
markers
ethoxyresorufin­
O­
dealkylase
(
EROD)
and
pentoxyresorufin­
O­
dealkylase
(
PROD),
were
observed
at
all
doses.
The
effect,
however,
was
not
dose­
related.
No
effect
on
the
CYP2E1
marker
pNP­
hydroxylase,
was
observed.
Based
on
kidney
and
liver
lesions
observed
at
the
mid
dose,
this
study
identified
a
NOAEL
of
75
mg/
kg­
day
and
a
LOAEL
of
150
mg/
kg­
day.

Thornton­
Manning
et
al.
(
1994)
conducted
an
analogous
experiment
with
female
C57BL/
6J
mice.
Six
mice
per
group
were
administered
an
aqueous
solution
(
10%
Emulphor
®
)
of
bromodichloromethane
by
gavage
for
five
consecutive
days
at
dose
levels
of
0,
75
and
150
mg/
kg­
day.
Animals
were
sacrificed
on
day
6.
All
mice
survived
to
the
termination
of
the
experiment.
No
effect
on
body,
kidney,
or
liver
weight
was
observed
with
the
exception
of
a
significant
increase
in
absolute
liver
weight
at
150
mg/
kg­
day.
No
change
in
cytochrome
P450
activity
was
observed,
although
a
nonsignificant
dose­
related
decrease
in
total
P450
content
was
observed.
ALT
was
significantly
increased
at
150
mg/
kg­
day,
and
a
significant
dose­
related
increase
in
SDH
activity
was
observed.
Creatinine
and
BUN
were
not
significantly
increased.
No
kidney
or
liver
lesions
were
observed
at
either
dose.
Based
on
increases
in
serum
enzyme
activity,
a
LOAEL
of
150
mg/
kg­
day
and
a
NOAEL
of
75
mg/
kg­
day
were
identified
for
this
study.
Draft
­
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February
20,
2003
V
­
17
Potter
et
al.
(
1996)
investigated
hyaline
droplet
formation
and
cell
proliferation
in
the
kidney
of
male
F344
rats.
Test
animals
(
4/
dose)
received
0.75
or
1.5
mmol/
kg
of
bromodichloromethane
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days.
The
administered
doses
corresponded
to
123
or
246
mg/
kg­
day.
No
exposure­
related
increase
in
hyaline
droplet
formation
was
observed
at
either
dose.
Binding
of
bromodichloromethane
to
 2u­
globulin
was
not
measured.
Cell
proliferation
in
the
kidney
was
assessed
in
vivo
by
[
3H]­
thymidine
incorporation.
No
statistically
significant
effect
of
bromodichloromethane
on
tubular
cell
proliferation
was
observed
following
exposures
of
up
to
7
days,
although
high
labeling
levels
were
observed
in
3
of
4
rats
at
the
246
mg/
kg­
day
dose.

Melnick
et
al.
(
1998)
exposed
female
B6C3F
1
mice
(
10
animals/
group)
to
bromodichloromethane
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
Doses
of
bromodichloromethane
used
in
this
study
were
0
(
vehicle
only),
75,
150,
or
326
mg/
kg­
day.
There
were
no
treatment
related
signs
of
overt
toxicity
observed
during
the
study.
Body
weight
and
water
intake
were
not
significantly
altered
at
any
dose
tested.
However,
a
significant
doserelated
increase
in
absolute
liver
weight
and
liver
weight/
body
weight
ratio
was
noted
for
the
150
and
326
mg/
kg­
day
dose
groups.
Serum
ALT
activity
was
significantly
increased
in
the
two
highest
dose
groups
and
serum
SDH
activity
was
elevated
at
all
doses
tested.
At
necropsy,
there
was
clear
evidence
of
hepatocyte
hydropic
degeneration
in
animals
treated
with
150
and
326
mg/
kg­
day.
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
The
two
highest
(
150
and
326
mg/
kgday
doses
resulted
in
significantly
elevated
hepatocyte
proliferation
as
measured
by
the
LI.
NOAEL
and
LOAEL
values
of
75
and
150
mg/
kg­
day
were
identified
on
the
basis
of
elevated
serum
enzyme
activity,
increased
liver
weight,
and
histological
findings.

NTP
(
1998)
evaluated
the
effect
of
bromodichloromethane
on
food
and
water
consumption
by
Sprague­
Dawley
rats
in
the
course
of
a
range­
finding
experiment
for
a
study
of
developmental
and
reproductive
effects.
This
study
was
conducted
in
compliance
with
the
Good
Laboratory
Practice
Regulations
as
described
in
21
CFR
58.
Bromodichloromethane
was
administered
to
test
animals
(
6
animals/
sex/
dose)
at
nominal
concentrations
of
0,
100,
500,
1000,
and
1500
ppm
in
the
drinking
water
for
2
weeks.
The
average
doses
of
bromodichloromethane
estimated
based
on
water
consumption
were
11,
45,
91
and
124
mg/
kg­
day
for
the
100,
500,
1000
and
1500
ppm
dose
groups,
respectively.
All
animals
were
observed
twice
daily
for
signs
of
toxicity.
Body
weight
data
were
obtained
twice
weekly
and
at
termination
of
the
experiment.
Feed
and
water
consumption
were
measured
twice
weekly.
Animals
were
euthanized
at
termination
of
the
experiment
without
necropsy.
No
mortality
or
treatment­
related
clinical
signs
were
observed
in
any
dose
group.
Body
weights
and
weight
gains
were
comparable
among
all
dose
groups,
except
for
body
weight
gains
on
Study
Day
5
(
the
first
day
of
compound
administration)
in
the
1000
and
1500
ppm
dose
groups
which
were
decreased
127.5%
and
118.5%,
respectively.
Feed
consumption
was
also
comparable
across
dose
groups,
with
the
exception
of
male
rats
dosed
with
1000
and
1500
ppm.
Male
rats
in
these
dose
groups
showed
decreases
in
consumption
of
31%
and
41%
,
respectively,
on
Study
Days
1
to
5.
Water
consumption
was
reduced
in
the
500,
1000,
and
1500
ppm
dose
groups,
suggesting
that
bromodichloromethane
is
unpalatable
at
higher
concentrations.
The
greatest
reduction
in
water
Draft
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February
20,
2003
V
­
18
intake
was
noted
on
Study
Days
1
to
5
(
61%
and
62%
for
males
in
the
1000
ppm
and
1500
ppm
dose
groups,
respectively,
and
38%,
40%
and
52%
for
females
in
the
500,
1000
and
1500
ppm
dose
groups,
respectively).

NTP
(
1998)
conducted
a
short­
term
reproductive
and
developmental
toxicity
screen
in
Sprague­
Dawley
rats
to
evaluate
the
potential
toxicity
of
bromodichloromethane
administered
in
drinking
water
for
35
days.
This
study
was
conducted
in
compliance
with
the
Good
Laboratory
Practice
Regulations
as
described
in
21
CFR
58.
Groups
of
male
and
female
rats
(
5­
13/
sex/
dose)
were
exposed
to
drinking
water
concentrations
of
0,
100,
700
and
1300
ppm
bromodichloromethane
using
the
study
design
described
in
Table
V­
6
(
Section
V.
E.
1).
Feed
and
water
consumption,
body
weight,
hematology,
clinical
chemistry,
cell
proliferation,
and
pathology
were
evaluated
in
addition
to
developmental
and
reproductive
endpoints.
Based
on
water
consumption
and
analytical
measurements
of
bromodichloromethane
in
the
provided
drinking
water,
the
calculated
average
daily
doses
were
0,
9,
38,
and
67
mg/
kg­
day
for
Group
A
males;
0,
7,
43,
and
69
mg/
kg­
day
for
Group
B
males;
and
0,
14,
69,
or
126
mg/
kg­
day
for
Group
C
females.

The
results
for
reproductive
and
developmental
effects
are
reported
in
Section
V.
E.
1.
Alterations
in
hematological
endpoints
or
clinical
chemistry
were
not
observed
following
bromodichloromethane
exposure,
with
the
exception
of
a
14%
drop
in
creatinine
in
the
100
ppm
Group
A
males
and
a
43%
increase
in
5­
nucleotidase
in
the
1300
ppm
Group
A
males
when
compared
to
controls.
An
increase
in
5­
nucleotidase
is
an
indication
of
hepatobiliary
dysfunction
in
which
there
is
interference
with
the
secretion
of
bile,
and
should
be
accompanied
by
a
parallel
change
in
alkaline
phosphatase
activity.
Since
alkaline
phophatase
activity
was
unaltered
in
this
study,
the
toxicological
significance
of
the
observed
increase
in
5­
nucleotidase
was
considered
uncertain.
Organ
weight
and
organ/
body
weight
ratios
reported
by
NTP
(
1998)
were
comparable
in
all
treatment
groups
for
both
males
and
females.
Histopathological
examination
identified
three
tissue
changes
that
were
potentially
treatment­
related.
Cytoplasmic
vacuolization
of
hepatocytes
and
mild
liver
necrosis
were
observed
in
Group
A
males
(
see
Table
V­
6
for
details
of
group
assignment)
treated
with
700
and
1300
ppm
bromodichloromethane
and
in
Group
B
males
treated
with
1300
ppm
bromodichloromethane.
Hepatic
necrosis
was
dose­
dependent,
with
incidences
of
0/
10,
0/
10,
4/
9,
and
10/
10
observed
at
0,
100,
700,
and
1300
ppm,
respectively.
These
changes
were
not
accompanied
by
an
increase
in
alkaline
phosphatase
activity.
Hematopoietic
cell
proliferation
in
the
spleen
was
observed
in
Group
A
males
at
all
doses
of
bromodichloromethane.
However,
the
biological
significance
of
this
finding
with
respect
to
bromodichloromethane
treatment
was
unclear,
since
cell
proliferation
in
the
spleen
may
occur
as
a
response
to
general
stress.
Evidence
of
mild
kidney
necrosis
was
evident
in
Group
A
males
in
the
1300
ppm
dose
group,
but
may
have
resulted
from
decreased
water
intake.
BrdU
labeling
index
(
LI),
a
measurement
of
cell
proliferation,
was
unchanged
in
the
livers
and
kidneys
of
Group
B
males
in
all
dose
groups.
A
small
but
statistically
significant
increase
in
the
LI
was
noted
in
the
livers
and
kidneys
of
Group
C
females
in
the
1300
ppm
dose
group.

As
discussed
in
Sections
V.
E.
1,
results
from
this
study
indicate
that
bromodichloromethane
did
not
result
in
reproductive
or
developmental
toxicity
at
drinking
water
concentrations
Draft
­
Do
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or
Quote
February
20,
2003
V
­
19
up
to
1300
ppm.
However,
exposure
to
concentrations
of
700
ppm
and
1300
ppm
produced
changes
in
liver
histopathology
in
male
rats
and
resulted
in
decreases
in
body
weight
and
food
and
water
consumption
in
both
sexes.
On
the
basis
of
these
results,
NTP
(
1998)
concluded
that
bromodichloromethane
is
unpalatable
at
these
concentrations
and
is
a
possible
general
toxicant
in
male
and
female
rats
at
concentrations
of
700
ppm
and
above.
Although
not
accompanied
by
changes
in
alkaline
phosphatase
activity,
the
occurrence
of
individual
hepatocyte
cell
necrosis
was
clearly
dose­
related
and
thus
considered
appropriate
for
identification
of
NOAEL
and
LOAEL
values.
Based
on
calculated
average
daily
doses
for
Group
A
males
at
the
100
and
700
ppm
concentrations
(
Table
6A
in
NTP,
1998),
these
data
identify
NOAEL
and
LOAEL
values
of
9
mg/
kg­
day
and
38
mg/
kg­
day,
respectively,
for
occurrence
of
hepatic
cell
necrosis.

Coffin
et
al.
(
2000)
examined
the
effect
of
bromodichloromethane
administered
by
gavage
in
corn
oil
or
in
drinking
water
on
cell
proliferation
and
DNA
methylation
in
the
liver
of
female
B6C3F1
mice.
Gavage
doses
of
0,
0.92,
or
1.83
mmol/
kg
(
0,
150,
or
300
mg/
kg,
respectively)
were
administered
to
test
animals
(
7­
8
weeks
old;
10/
group)
daily
for
five
days,
off
for
two
days,
and
then
again
daily
for
four
days.
The
high
dose
was
selected
on
the
basis
that
it
had
previously
been
shown
to
be
carcinogenic
in
female
mice.
In
a
separate
experiment,
bromodichloromethane
was
administered
in
drinking
water
for
11
days
at
approximately
75%
of
the
saturation
level,
resulting
in
an
average
daily
dose
of
0.85
mmol/
kg
(
138
mg/
kg).
The
mice
were
sacrificed
24
hours
after
the
last
gavage
dose
and
the
livers
were
removed,
weighed,
and
processed
for
histopathological
examination,
proliferating
cell
nuclear
antigen
­
labeling
index
(
PCNA­
LI)
analysis,
and
determination
of
c­
myc
methylation
status.
A
significant,
dose­
dependent
increase
in
relative
liver
weight
was
observed
in
animals
dosed
by
gavage;
however,
relative
liver
weight
was
unaffected
in
animals
administered
the
compound
in
drinking
water,
when
compared
to
controls.
Histopathological
findings
in
gavage­
dosed
animals
consisted
of
hydropic
degeneration
at
the
low
dose
and
necrosis,
fibrosis,
and
giant
cell
reaction
at
the
high
dose.
No
severity
or
incidence
data
were
provided.
The
histopathology
findings
for
animals
receiving
bromodichloromethane
in
the
drinking
water
were
similar
to
those
observed
in
the
low
dose
gavage
group.
Bromodichloromethane
administered
by
gavage
caused
a
dose­
dependent
increase
in
the
PCNA­
LI
which
was
significant
at
each
dose
tested
when
compared
to
the
control.
There
was
no
significant
effect
when
the
compound
was
administered
in
drinking
water.
Administration
of
bromodichloromethane
by
gavage
or
in
drinking
water
decreased
methylation
of
the
c­
myc
gene.
A
LOAEL
of
150
mg/
kg,
the
lowest
dose
tested,
was
identified
on
the
basis
of
liver
toxicity
(
hydropic
degeneration)
and
increased
cell
proliferation
in
animals
administered
bromodichloromethane
by
corn
oil
gavage.
A
NOAEL
was
not
identified.
The
results
of
the
single­
dose
drinking
water
experiment
suggest
a
slightly
lower
LOAEL
of
138
mg/
kg­
day,
based
on
hydropic
degeneration
of
the
liver.

Torti
et
al.
(
2001)
conducted
a
one
week
inhalation
exposure
study
of
bromodichloromethane
in
male
wild­
type
(
p53+/+)
and
genetically
engineered
p53
heterozygous
(
p53+/­)
mice.
C57BL/
6,
FVB/
N,
and
C57BL/
6
p53+/­
mice
(
6
mice/
type/
concentration)
were
exposed
to
target
exposure
concentrations
of
0,
1,
10,
30,
100,
or
150
ppm
for
six
hours
per
day,
seven
days
per
week.
FVB/
N
p53+/­
mice
were
exposed
to
concentrations
of
0,
0.3,
1,
10,
or
30
ppm
for
six
hours
per
day,
seven
days
per
week.
The
test
animals
were
evaluated
for
clinical
and
Draft
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or
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February
20,
2003
V
­
20
pathological
changes
and
induced
regenerative
cell
proliferation
in
kidney
and
liver.
Osmotic
pumps
for
delivery
of
bromodeoxyuridine
for
determination
of
labeling
index
were
implanted
at
3.5
days
prior
to
scheduled
termination.
Test
animals
were
euthanized
approximately
18
hours
after
the
last
scheduled
exposure.
Average
measured
concentrations
of
bromodichloromethane
in
the
middle
dose
range
were
102
to
104%
of
the
target
concentration
(
coefficients
of
variation
2.6
to
10.6%).
The
lowest
dose
average
concentration
was
114%
(
coefficient
of
variation
38.6%).
The
average
high
dose
concentration
was
78.8%
of
the
target
concentration
as
a
result
of
technical
problems
with
the
metering
system.
Deaths
were
observed
at
concentrations
of
30
ppm
and
greater,
with
100%
mortality
observed
for
C57BL/
6
heterozygous
and
FVB/
N
wild
type
mice
at
150
ppm.
Clinical
signs
in
mice
surviving
exposure
at
100
and
150
ppm
included
lethargy
and
labored
breathing.
Reddened
skin
and
eyes
were
also
observed,
primarily
at
concentrations
of
30
ppm
and
above.
Relative
body
weight
was
decreased
at
exposure
concentrations
of
30
ppm
and
above.
Body
weight
loss
was
greater
in
the
heterozygous
p53+/­
than
in
the
corresponding
wild
type
strain.
Significantly
increased
relative
kidney
weight
was
evident
in
all
mice
exposed
to
concentrations
of
30
ppm
and
above
and
in
FVB/
N
mice
exposed
to
10
ppm.
Significantly
increased
relative
liver
weight
was
observed
in
wild
type
and
heterozygous
FVB/
N
mice
exposed
to
concentrations
of
10
ppm
and
above.
Reduced
body
weight
gains
were
observed
in
surviving
mice
exposed
to
30
ppm
and
above,
with
greater
reductions
observed
in
heterozygous
mice
when
compared
to
wild
type
mice.
Histopathologic
evaluation
revealed
severe
renal
damage
consisting
of
nephrosis,
tubular
degeneration,
and
associated
regeneration.
The
averaged
severity
scores
indicated
greater
damage
in
heterozygous
compared
to
wild
type
mice
and
in
FVB/
N
compared
to
the
C57BL/
N
mice.
Centrilobular
degeneration
and
necrosis
were
observed
in
the
livers
of
moribund
mice
sacrificed
before
study
termination.
These
lesions
were
also
observed
in
surviving
animals
at
study
termination.
Hydropic
degeneration
occurred
at
concentrations
of
30
ppm
and
above
in
wild
type
and
heterozygous
C57BL/
N
mice
and
at
concentrations
of
10
ppm
and
above
in
FVB/
N
mice.
Necrosis
was
evident
at
concentrations
of
100
and
150
ppm.
No
histopathologic
lesions
were
observed
in
the
bladder.
Regenerative
cell­
proliferation
in
the
kidney
cortex
was
significantly
increased
at
exposure
concentrations
of
10
ppm
and
above.
Renal
labeling
indices
were
approximately
20
to
30
%
at
10
ppm
and
45
to
60%
at
the
high
dose.
Regenerative
cell
proliferation
in
the
liver
was
less
pronounced
than
in
the
kidney.
Minimal
increases
in
the
labeling
index
(
approximately
10%
or
less)
were
observed
in
wild
type
C57BL/
N
and
FVB/
N
mice.
Labeling
index
was
significantly
increased
only
in
C57BL/
N
wild
type
mice
exposed
at
the
100
ppm
level.
A
modest
increase
in
labeling
index
was
observed
in
heterozygous
FVB/
N
mice
exposed
to
10
or
30
ppm;
the
response
reached
statistical
significance
at
30
ppm.
Relatively
large
increases
in
labeling
index
(
up
to
40%)
were
observed
in
heterozygous
C57BL/
N
mice
at
doses
of
10
ppm
and
above.
The
response
was
statistically
significant
at
the
30
and
100
ppm
exposure
levels.
Bromodichloromethane
did
not
induce
cellular
proliferation
in
the
transitional
epithelium
of
the
bladder.
These
data
identify
NOAEL
and
LOAEL
values
of
1
and
10
ppm,
respectively,
based
on
histopathological
changes
in
the
kidney
of
male
p53
wild
type
and
heterozygous
C57BL/
6
and
FVB/
N
mice.

Torti
et
al.
(
2001)
also
conducted
a
three
week
inhalation
exposure
study
of
bromodichloromethane
in
wild­
type
(
p53+/+)
and
genetically
engineered
p53
heterozygous
(
p53+/­)
male
mice.
C57BL/
6,
FVB/
N,
C57BL/
6
p53+/­,
and
FVB/
N
p53+/­
mice
(
6
Draft
­
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or
Quote
February
20,
2003
V
­
21
mice/
type/
concentration)
were
exposed
to
target
exposure
concentrations
of
0.3,
1,
3,
10,
or
30
ppm
for
six
hours
per
day,
seven
days
per
week.
The
test
protocol
and
endpoints
measured
were
the
same
as
those
used
for
the
one
week
study
described
above.
Test
animals
were
euthanized
approximately
18
hours
after
the
last
scheduled
exposure.
Average
measured
concentrations
were
92
to
97%
of
the
target
concentrations,
with
coefficients
of
variation
ranging
from
3.1
to
6.9%.
Mortality
was
observed
in
all
30
ppm
dose
groups
with
the
exception
of
wild
type
C57BL/
6
mice.
No
clinical
signs
of
toxicity
were
reported.
Body
weight
gain
was
significantly
reduced
only
in
C57BL/
6
wild
type
mice
exposed
at
30
ppm.
Relative
kidney
weights
in
exposed
groups
did
not
differ
significantly
from
the
control
values.
Significantly
increased
relative
liver
weight
was
observed
in
only
in
heterozygous
C57BL/
6
and
wild
type
FVB/
N
mice
exposed
at
30
ppm.
Histopathologic
evaluation
revealed
near­
normal
kidney
architecture.
Minimal
to
moderate
degenerative
tubular
change
and
regenerative
tubules
were
observed
in
the
10
and
30
ppm
groups,
but
the
acute
tubular
nephrosis
observed
in
the
one
week
study
was
not
evident.
Minimal
hepatocyte
degeneration
was
observed
in
heterozygous
C57BL/
6
mice
exposed
at
30
ppm
and
in
heterozygous
FVB/
N
mice
exposed
at
10
or
30
ppm.
These
observations
suggest
that
the
liver
and
severe
renal
toxicity
observed
in
the
one
week
experiment
conducted
by
Torti
et
al.
(
2001)
are
transient
and
were
resolving
by
three
weeks.
No
histopathologic
lesions
were
observed
in
the
bladder.
Regenerative
cell­
proliferation
in
the
kidney
cortex
was
near
baseline
levels,
with
only
the
30
ppm
groups
showing
small
elevations.
These
elevations
were
statistically
significant
in
all
30
ppm
groups
except
C57BL/
N
wild
type
mice.
No
increases
in
regenerative
cell
proliferation
were
evident
in
the
liver
or
bladder.
The
NOAEL
and
LOAEL
values
in
this
study
are
3
and
10
ppm,
respectively,
based
on
histopathologic
changes
in
the
liver
and
kidney
of
male
p53
wild
type
and
heterozygous
C57BL/
6
and
FVB/
N
mice.

2.
Dibromochloromethane
Chu
et
al.
(
1982a)
administered
dibromochloromethane
to
male
Sprague­
Dawley
rats
(
10/
group)
in
drinking
water
for
28
days
at
dose
levels
of
0,
5,
50,
or
500
ppm.
Based
on
recorded
fluid
intake,
these
levels
corresponded
to
doses
of
0,
0.7,
8.5,
or
68
mg/
kg­
day,
as
calculated
by
the
authors.
The
authors
observed
no
effects
on
growth
rate
or
food
consumption
and
no
signs
of
toxicity
throughout
the
exposure.
No
dose­
related
biochemical
or
histologic
changes
were
detected
(
no
data
provided).
This
study
identified
a
NOAEL
of
68
mg/
kg­
day,
but
the
reported
data
were
too
limited
to
allow
an
independent
verification.

Munson
et
al.
(
1982)
administered
dibromochloromethane
by
aqueous
gavage
to
male
and
female
CD­
1
mice
(
8
to
12/
sex/
group)
for
14
days
at
dose
levels
of
0,
50,
125,
or
250
mg/
kg­
day.
Endpoints
measured
included
body
and
organ
weights,
hematology,
clinical
chemistry,
and
humoral
and
cell­
mediated
immune
system
function.
At
250
mg/
kg­
day,
body
weights
were
significantly
decreased
only
in
high­
dose
males.
Significant
organ
weight
changes
included
increased
absolute
liver
weight
(
high­
dose
females),
increased
relative
liver
weight
(
mid­
and
highdose
groups),
and
decreased
absolute
and
relative
spleen
weight
(
high­
dose
males).
The
only
hematology
parameter
significantly
affected
by
treatment
was
fibrinogen
concentration,
which
was
decreased
in
high­
dose
males
and
females.
Significant
clinical
chemistry
findings
were
limited
to
the
high­
dose
groups.
Specifically,
glucose
levels
were
significantly
decreased
in
both
males
and
Draft
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February
20,
2003
V
­
22
females,
and
ALT
and
AST
activities
were
significantly
increased
in
both
males
and
females.
Dibromochloromethane
appeared
to
affect
the
humoral
immune
system,
as
judged
by
significantly
decreased
antibody­
forming
cells
(
mid­
and
high­
dose
groups)
and
hemagglutination
titers
(
highdose
groups).
The
cell­
mediated
immune
system
also
appeared
to
be
affected
in
male
animals,
as
judged
by
a
significant
decrease
in
the
popliteal
lymph
node
stimulation
index
at
the
high
dose.
This
study
identified
a
NOAEL
of
50
mg/
kg­
day
and
a
LOAEL
of
125
mg/
kg­
day
for
dibromochloromethane,
based
on
decreased
immune
function.

Condie
et
al.
(
1983)
investigated
the
renal
and
hepatic
toxicity
of
dibromochloromethane.
Male
CD­
1
mice
(
8
to
16/
group)
were
administered
0,
37,
74,
or
147
mg/
kg­
day
of
dibromochloromethane
by
gavage
in
corn
oil
for
14
days.
Biochemical
evidence
of
liver
damage
(
elevated
ALT)
and
kidney
damage
(
decreased
PAH
uptake
by
kidney
slices)
was
observed
at
the
high
dose,
but
not
at
the
mid­
level
or
low
doses.
Similarly,
histopathology
revealed
no
consistent
or
important
changes
at
the
low
or
mid
doses
with
minimal
to
moderate
liver
and
kidney
injury
at
the
high
dose.
Liver
lesions
included
mitotic
figures,
focal
inflammation,
and
cytoplasmic
vacuolation,
while
kidney
lesions
included
epithelial
hyperplasia
and
mesangial
nephrosis.
On
this
basis,
this
study
identified
a
NOAEL
value
of
74
mg/
kg­
day
and
LOAEL
value
of
147
mg/
kg­
day
for
dibromochloromethane.

In
a
14­
day
study
by
NTP
(
1985),
groups
of
male
and
female
F344/
N
rats
(
5/
sex/
dose)
were
administered
0,
60,
125,
250,
500,
or
1,000
mg/
kg­
day
of
dibromochloromethane
by
gavage
in
corn
oil.
Animals
were
observed
twice
daily
for
mortality
and
were
weighed
once
per
week.
Necropsies
were
performed
on
all
animals.
All
high­
dose
rats
and
all
females
that
received
500
mg/
kg­
day
died
by
day
6.
Three
males
at
500
mg/
kg­
day
died
between
days
5
and
8.
No
deaths
occurred
at
or
below
250
mg/
kg­
day.
At
500
or
1000
mg/
kg­
day,
clinical
observations
included
lethargy,
ataxia,
and
labored
breathing.
Treatment­
related
macroscopic
findings
included
mottled
livers
and
darkened
renal
medullae
in
animals
administered
500
or
1,000
mg/
kg­
day.
Based
on
behavior,
gross
pathology,
and
mortality,
this
study
identified
a
NOAEL
of
250
mg/
kg­
day
and
a
LOAEL
of
500
mg/
kg­
day.

In
a
parallel
study
(
NTP,
1985),
male
and
female
B6C3F
1
mice
(
5/
sex/
dose)
were
administered
0,
30,
60,
125,
250,
or
500
mg/
kg­
day
of
dibromochloromethane
in
corn
oil
by
gavage
for
14
days.
Treatment­
related
deaths
occurred
in
80%
of
the
males
and
in
60%
of
the
females
at
the
high
dose.
Clinical
signs
at
this
dose
included
lethargy,
ataxia,
and
labored
breathing.
Treatment­
related
macroscopic
findings
included
mottled
livers
and
darkened
renal
medullae
in
high­
dose
males
and
females.
White
papillomatous
nodules
in
the
stomach
were
also
observed
in
males
at
125,
250,
or
500
mg/
kg­
day
and
in
female
mice
at
250
or
500
mg/
kg­
day.
Based
on
gross
lesions,
this
study
identified
a
NOAEL
of
60
mg/
kg­
day
and
a
LOAEL
of
125
mg/
kg­
day
in
mice.

Aida
et
al.
(
1992a)
investigated
the
effects
of
administering
dibromochloromethane
to
Slc:
Wistar
rats
(
7/
sex/
group)
for
one
month
at
dietary
levels
of
0%,
0.020%,
0.062%,
or
0.185%
for
males,
and
0%,
0.038%,
0.113%,
or
0.338%
for
females.
The
test
material
was
microencapsulated
and
mixed
with
powdered
feed;
placebo
granules
were
used
for
the
control
Draft
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February
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2003
V
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23
groups.
Based
on
the
mean
food
intakes,
the
study
authors
reported
calculated
doses
of
0,
18.3,
56.2,
or
173.3
mg/
kg­
day
for
males
and
0,
34.0,
101.1,
or
332.5
mg/
kg­
day
for
females.
Clinical
effects,
body
weight,
food
consumption,
hematology
parameters,
serum
chemistry,
and
histopathology
of
all
major
organs
were
determined.
Body
weights
were
significantly
reduced
in
high­
dose
females
relative
to
the
controls.
High­
dose
females
also
exhibited
slight
piloerection
and
emaciation.
Dose­
related
increases
in
both
absolute
and
relative
liver
weights
were
observed
in
males
(
significant
at
the
high
dose)
and
females
(
significant
at
all
dose
levels
with
the
exception
of
a
nonsignificant
increase
in
absolute
liver
weight
at
the
low
dose).
Relative
kidney
weights
were
also
significantly
increased
in
the
high­
dose
females.
Significant
decreases
in
alkaline
phosphatase
(
mid­
and
high­
dose
males
and
all
female
dose
groups)
and
LDH
(
all
female
dose
groups)
were
observed,
but
the
biological
significance
of
these
changes
is
unclear.
Significant,
dose­
related
changes
in
serum
biochemistry
included
reduced
nonesterified
fatty
acids
in
highdose
males,
reduced
T­
GLY
in
high­
dose
groups,
and
increased
cholesterol
in
mid­
and
high­
dose
males
and
in
females
at
all
dose
levels.
The
cholesterol
levels,
however,
were
within
normal
ranges
at
all
dose
levels.
Serum
cholinesterase
activity
was
also
significantly
decreased
in
highdose
males
and
mid­
and
high­
dose
females
with
the
trend
clearly
dose­
related
in
females.
Liver
cell
vacuolization
was
generally
noted
at
a
similar
incidence
in
the
controls
and
all
dosing
groups,
but
dose­
related
increases
in
severity
were
observed
in
mid­
and
high­
dose
males
and
females.
The
incidence
and
very
slight
severity
of
the
effects
at
the
low
dose
were
similar
to
those
observed
in
the
control
groups
and
were
not
considered
adverse.
The
severity
of
the
liver
cell
vacuolization
at
the
mid­
dose
was
rated
as
very
slight
to
slight,
while
the
severity
at
the
high­
dose
was
rated
as
moderate
to
remarkable.
Swelling
and
single
cell
necrosis
were
also
observed,
primarily
in
the
high­
dose
groups.
No
effect
was
observed
on
any
hematology
parameter.
Based
on
the
histopathology
findings,
NOAELs
of
18.3
(
males)
and
34.0
(
females)
mg/
kg­
day
and
LOAELs
of
56.2
(
males)
and
101.1
(
females)
mg/
kg­
day
were
identified
for
dibromochloromethane
in
rats.

Potter
et
al.
(
1996)
evaluated
hyaline
droplet
formation
and
cell
proliferation
in
the
kidney
of
male
F344
rats
following
exposure
to
dibromochloromethane.
The
rats
(
4/
dose)
were
dosed
with
0.75
or
1.5
mmol/
kg
(
156
or
312
mg/
kg­
day,
respectively)
of
dibromochloromethane
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days.
No
exposure­
related
increase
in
hyaline
droplets
was
observed
in
dosed
rats.
Binding
to
 2u­
globulin
was
not
measured.
Changes
in
kidney
tubule
cell
proliferation
were
assessed
by
in
vivo
incorporation
of
[
3H]­
thymidine.
No
statistically
significant
effect
of
dibromochloromethane
exposure
on
this
endpoint
was
noted
following
exposures
of
up
to
7
days
duration.

Melnick
et
al.
(
1998)
exposed
female
B6C3F
1
mice
(
10/
dose)
to
dibromochloromethane
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
The
doses
of
dibromochloromethane
in
this
study
were
0
(
vehicle
only),
50,
100,
192,
or
417
mg/
kg­
day.
The
corresponding
time­
weighted
doses
were
0,
37,
71,
137,
and
298
mg/
kg­
day.
No
treatment­
related
signs
of
overt
toxicity
were
observed
during
the
study.
Body
weight
and
water
intake
were
not
significantly
altered
at
any
dose
tested.
However,
a
statistically
significant
and
dose­
related
increase
in
liver
weight/
body
weight
ratio
was
seen
in
the
100,
192
and
417
mg/
kg­
day
dose
groups.
Serum
ALT
activity
was
significantly
increased
in
the
two
highest
dose
groups.
The
activity
of
serum
SDH
was
Draft
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or
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February
20,
2003
V
­
24
significantly
elevated
at
all
doses
tested
except
50
mg/
kg­
day.
However,
the
increase
in
activity
(
shown
graphically)
was
very
small
relative
to
the
control
at
the
100
and
192
mg/
kg­
day
doses.
At
necropsy,
there
was
clear
evidence
of
hepatocyte
hydropic
degeneration
in
the
192
and
417
mg/
kg­
day
dose
groups.
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
Only
the
highest
dose
tested
(
417
mg/
kg­
day)
resulted
in
significantly
elevated
hepatocyte
proliferation
as
measured
by
the
LI.
Evaluation
of
the
data
in
this
study
suggest
a
LOAEL
of
192
mg/
kg­
day,
based
on
a
consistent
pattern
of
positive
results
for
indicators
of
hepatotoxicity
at
this
dose.

Coffin
et
al.
(
2000)
examined
the
effect
of
dibromochloromethane
administered
by
gavage
in
corn
oil
or
in
drinking
water
on
cell
proliferation
and
DNA
methylation
in
the
liver
of
female
B6C3F1
mice.
Gavage
doses
of
0,
0.48,
or
1.44
mmol/
kg
(
0,
100,
or
300
mg/
kg,
respectively)
were
administered
to
test
animals
(
7­
8
weeks
old;
10/
group)
daily
for
five
days,
off
for
two
days,
and
then
again
daily
for
four
days.
The
high
dose
was
selected
on
the
basis
that
it
had
previously
been
demonstrated
to
be
carcinogenic
in
female
mice.
Dibromochloromethane
was
administered
in
drinking
water
at
approximately
75%
of
the
saturation
level,
resulting
in
an
average
daily
dose
of
0.82
mmol/
kg
(
171
mg/
kg).
The
mice
were
sacrificed
24
hours
after
the
last
gavage
dose
and
the
livers
were
removed,
weighed,
and
processed
for
histopathological
examination,
proliferating
cell
nuclear
antigen
­
labeling
index
(
PCNA­
LI)
analysis,
and
determination
of
c­
myc
methylation
status.
For
histopathological
analysis,
stained
liver
sections
were
evaluated
for
toxicity
using
a
semi­
quantitative
procedure
using
the
following
severity
scoring
system:
Grade
1
consisted
of
mid
lobular
ballooning
hepatocytes;
Grade
2
consisted
of
mid
lobular
ballooning
hepatocytes
extending
to
the
central
vein;
Grade
3
consisted
of
centrilobular
necrosis
with
ballooning
hepatocytes;
and
Grade
4
consisted
of
necrosis
extending
from
the
central
vein
to
the
mid
lobule
zone.
A
significant,
dose­
dependent
increase
in
relative
liver
weight
was
observed
in
animals
dosed
by
gavage;
however,
relative
liver
weight
was
unaffected
in
animals
administered
the
compound
in
drinking
water,
when
compared
to
controls.
At
the
low
gavage
dose,
liver
toxicity
consisted
mainly
of
a
Grade
1
response.
At
the
high
dose,
liver
toxicity
consisted
mainly
of
a
Grade
2
response.
No
incidence
data
were
provided
in
the
study
report,
nor
was
a
severity
grade
reported
for
the
control
group.
The
histopathology
findings
for
animals
receiving
bromodichloromethane
in
the
drinking
water
were
similar
to
those
observed
in
the
low
dose
gavage
group.
Dibromochloromethane
administered
by
gavage
caused
a
dose­
dependent
increase
in
the
PCNA­
LI.
The
increases
observes
at
each
dose
were
significantly
different
from
the
control.
There
was
no
significant
effect
on
PCNA­
LI
when
the
compound
was
administered
in
drinking
water.
Administration
of
dibromochloromethane
by
gavage
or
in
drinking
water
decreased
methylation
of
the
c­
myc
gene.
A
LOAEL
of
100
mg/
kg,
the
lowest
dose
tested,
was
identified
on
the
basis
of
liver
toxicity
(
ballooning
hepatocytes)
and
increased
cell
proliferation
in
gavaged
animals.

3.
Bromoform
Chu
et
al.
(
1982a)
administered
bromoform
to
male
Sprague­
Dawley
rats
(
10/
group)
in
drinking
water
for
28
days
at
dose
levels
of
0,
5,
50,
or
500
ppm.
Based
on
recorded
fluid
intake,
these
levels
corresponded
to
doses
of
0,
0.7,
8.5,
or
80
mg/
kg­
day,
as
calculated
by
the
authors.
Draft
­
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or
Quote
February
20,
2003
V
­
25
The
authors
observed
no
effects
on
growth
rate
or
food
consumption
and
no
signs
of
toxicity
throughout
the
exposure.
No
dose­
related
biochemical
or
histologic
changes
were
detected
(
no
data
provided).
This
study
identified
a
NOAEL
of
80
mg/
kg­
day,
but
the
reported
data
were
too
limited
to
allow
an
independent
confirmation.

Munson
et
al.
(
1982)
administered
bromoform
by
aqueous
gavage
to
male
and
female
CD­
1
mice
(
6
to
12/
sex/
group)
for
14
days
at
levels
of
0,
50,
125,
or
250
mg/
kg­
day.
Parameters
observed
included
body
and
organ
weights,
hematology,
clinical
chemistry,
and
humoral
and
cellmediated
immune
system
functions.
Body
weights
were
significantly
decreased
in
high­
dose
females,
while
body
weights
in
males
were
significantly
increased
at
the
mid
and
high
doses.
Absolute
and
relative
liver
weights
were
significantly
increased
in
males
at
the
mid
and
high
dose
and
in
females
at
the
high
dose.
Absolute
spleen
weight
was
also
decreased
in
mid­
and
high­
dose
females.
Hematologic
effects
included
significantly
decreased
fibrinogen
in
males
at
the
high
dose
and
significantly
decreased
prothrombin
time
in
all
treated
males.
The
changes
in
prothrombin
time,
however,
were
not
dose­
related.
Significant
clinical
chemistry
findings
included
decreased
glucose
levels
(
high­
dose
males),
increased
AST
activity
(
high­
dose
groups),
and
decreased
BUN
levels
(
high­
dose
males).
Both
the
humoral
and
cell­
mediated
immune
systems
appeared
to
be
affected
in
males
at
the
high
dose
with
a
significant
decrease
in
antibody­
forming
cells
and
a
significant
decrease
in
delayed­
type
hypersensitivity
response.
The
authors
stated
that
no
treatment­
related
effects
on
the
immune
system
in
females
were
observed
(
no
data
were
reported).
Based
on
changes
in
clinical
chemistry
parameters,
this
study
identified
a
NOAEL
of
125
mg/
kg­
day
and
a
LOAEL
of
250
mg/
kg­
day.

Condie
et
al.
(
1983)
investigated
the
renal
and
hepatic
toxicity
of
bromoform.
Male
CD­
1
mice
(
8
to
16/
group)
were
administered
0,
72,
145,
or
289
mg/
kg­
day
of
bromoform
by
gavage
in
corn
oil
for
14
days.
Biochemical
evidence
of
liver
damage
(
elevated
ALT)
and
kidney
damage
(
decreased
PAH
uptake
by
kidney
slices)
was
observed
at
the
high
dose,
but
not
at
the
mid
or
low
dose.
Histopathological
examination
revealed
no
consistent
or
important
changes
at
the
low
or
mid
doses,
with
minimal
to
moderate
liver
and
kidney
injury
at
the
high
dose.
Specific
microscopic
changes
included
intratubular
mineralization,
epithelial
hyperplasia,
mesangial
hypertrophy
and
mesangial
nephrosis
in
the
kidney,
and
centrilobular
pallor,
mitotic
figures,
focal
inflammation,
and
cytoplasmic
vacuolation
in
the
liver.
On
this
basis,
this
study
identified
a
NOAEL
value
of
145
mg/
kg­
day
and
a
LOAEL
value
of
289
mg/
kg­
day.

NTP
(
1989a)
investigated
the
short
term
oral
toxicity
of
bromoform
in
F344/
N
rats
and
B6C3F
1
mice.
Groups
of
male
and
female
rats
(
5/
sex/
group)
and
female
mice
(
5/
group)
were
administered
doses
of
0,
100,
200,
400,
600,
or
800
mg/
kg­
day
of
bromoform
in
corn
oil
by
gavage
for
14
days.
Male
mice
were
administered
0,
50,
100,
200,
400,
or
600
mg/
kg­
day.
All
rats
that
were
dosed
at
600
or
800
mg/
kg­
day
died
before
the
end
of
the
study.
At
400
mg/
kgday
only
one
male
rat
died
before
study
termination.
These
rats
exhibited
lethargy,
labored
breathing,
and
ataxia.
At
400
mg/
kg­
day,
final
body
weights
were
decreased
by
14%
in
male
rats
and
by
4%
in
female
rats
relative
to
controls.
In
mice,
one
male
and
one
female
administered
the
high
dose
died
before
study
termination.
At
dose
levels
of
600
mg/
kg­
day
or
above,
ataxia
and
lethargy
were
noted.
Final
body
weights
of
mice
were
comparable
to
those
of
the
controls.
Draft
­
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or
Quote
February
20,
2003
V
­
26
Raised
stomach
nodules
were
observed
in
males
at
400
and
600
mg/
kg­
day
and
in
females
at
600
and
800
mg/
kg­
day.
Based
on
decreased
body
weight
and
mortality
in
rats
and
on
stomach
nodules
in
mice,
this
study
identified
a
NOAEL
of
200
mg/
kg­
day
and
a
LOAEL
of
400
mg/
kgday

Aida
et
al.
(
1992a)
administered
bromoform
to
Slc:
Wistar
rats
(
7/
sex/
group)
for
one
month
at
dietary
levels
of
0%,
0.068%,
0.204%,
or
0.612%
for
males
and
0%,
0.072%,
0.217%,
or
0.651%
for
females.
The
test
material
was
microencapsulated
and
mixed
with
powdered
feed;
placebo
granules
were
used
for
the
control
groups.
Based
on
the
mean
food
intakes,
the
study
authors
reported
the
mean
compound
intakes
as
0,
61.9,
187.2,
or
617.9
mg/
kg­
day
for
males
and
0,
56.4,
207.5,
or
728.3
mg/
kg­
day
for
females.
Clinical
effects,
body
weight,
food
consumption,
hematology
parameters,
serum
chemistry,
and
histopathology
of
all
major
organs
were
determined.
Body
weights
were
significantly
reduced
in
high­
dose
males
relative
to
the
controls.
High­
dose
animals
of
both
sexes
exhibited
slight
piloerection
and
emaciation.
Relative
liver
weight
was
significantly
increased
in
mid­
and
high­
dose
males
and
females.
Significant
changes
in
serum
chemistry
were
primarily
observed
in
the
mid­
and
high­
dose
animals
with
the
females
more
significantly
affected.
These
changes
included
significant
decreases
in
(
a)
serum
glucose
in
low­
and
high­
dose
males
and
in
mid­
and
high­
dose
females,
(
b)
triglycerides
in
high­
dose
males
and
in
mid­
and
high­
dose
females,
(
c)
cholinesterase
activity
in
high­
dose
males
and
in
all
female
treatment
groups,
(
d)
LDH
in
mid­
and
high­
dose
females,
and
(
e)
BUN
in
mid­
and
high­
dose
females.
All
of
these
changes
in
the
groups
noted
exhibited
strong
dose­
related
trends
with
the
exception
of
serum
glucose
in
males.
Creatinine
levels
and
alkaline
phosphatase
activity
were
also
significantly
decreased
in
all
female
treatment
groups,
but
the
changes
were
not
dose­
related.
Significant
increases,
although
not
dose­
related,
were
observed
for
phospholipids
and
cholesterol
in
mid­
and
high­
dose
animals
with
the
exception
of
a
nonsignificant
increase
in
phospholipids
in
high­
dose
females.
The
only
change
of
clear
biological
significance
at
the
low
dose
was
a
decrease
in
cholinesterase
activity
in
females.
No
effect
was
observed
on
any
hematology
parameter.
Microscopic
and
macroscopic
findings
were
limited
to
the
liver.
Specifically,
discoloration
was
observed
in
all
males
and
females
in
the
high­
dose
group.
The
incidence
and
severity
of
liver
cell
vacuolization
and
swelling
were
dose­
related.
Severe
hepatic
cell
vacuolization
was
observed
in
5/
7
high­
dose
males
and
in
6/
7
females
at
the
mid
and
high
dose.
Slight
to
moderate
liver
cell
swelling
was
observed
in
three
high­
dose
males,
while
all
high­
dose
females
displayed
slight
signs
of
liver
cell
swelling.
Females
appeared
to
be
more
sensitive
for
development
of
histopathological
effects,
but
the
changes
observed
in
low­
dose
females
were
not
considered
an
adverse
effect.
Based
on
the
histopathology
and
serum
chemistry
changes
in
the
mid­
dose
animals,
this
study
identified
NOAELs
of
61.9
mg/
kg­
day
for
males
and
56.4
mg/
kg­
day
for
females,
and
LOAELs
of
187.2
mg/
kg­
day
for
males
and
207.5
mg/
kg­
day
for
females.

Potter
et
al.
(
1996)
evaluated
the
effect
of
bromoform
on
hyaline
droplet
formation
and
cell
proliferation
in
the
kidney
of
male
F344
rats.
Animals
(
4/
dose)
received
doses
of
0.75
or
1.5
mmol/
kg
of
bromoform
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days.
These
doses
correspond
to
190
or
379
mg/
kg­
day,
respectively.
No
exposure­
related
increase
in
hyaline
droplet
formation
was
observed.
Cell
proliferation
in
the
kidney
following
bromoform
exposure
was
measured
in
vivo
by
[
3H]­
thymidine
incorporation.
No
statistically
significant
effects
were
noted
Draft
­
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February
20,
2003
V
­
27
following
exposures
of
up
to
7
days
duration.
Binding
of
bromoform
to
 2u­
globulin
was
not
measured.

Melnick
et
al.
(
1998)
exposed
female
B6C3F
1
mice
(
10
animals/
group)
to
bromoform
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
Doses
of
bromoform
used
in
this
study
were
0
(
vehicle
only),
200,
or
500
mg/
kg­
day.
There
were
no
treatment­
related
signs
of
overt
toxicity
observed
during
the
study.
Body
weight
and
water
intake
were
not
significantly
altered
at
any
dose
tested.
However,
a
dose­
related
increase
in
absolute
liver
weight
and
liver
weight/
body
weight
ratio
was
noted
in
both
tested
doses.
Neither
serum
ALT
nor
serum
SDH
activity
were
significantly
elevated
at
either
dose
of
bromoform.
At
necropsy,
there
was
no
evidence
of
hepatocyte
hydropic
degeneration
in
animals
treated
with
either
dose.
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
Only
the
500
mg/
kg­
day
dose
resulted
in
marginally
significant
increase
in
hepatocyte
proliferation
as
measured
by
the
LI.
These
data
suggest
a
NOAEL
of
200
mg/
kg­
day
and
a
LOAEL
of
500
mg/
kg­
day
based
on
increased
hepatocyte
proliferation.

Coffin
et
al.
(
2000)
examined
the
effect
of
bromoform
administered
by
gavage
in
corn
oil
or
in
drinking
water
on
liver
toxicity,
cell
proliferation
and
DNA
methylation
in
female
B6C3F1
mice.
Gavage
doses
of
0,
0.79,
or
1.98
mmol/
kg
(
0,
200,
or
500
mg/
kg,
respectively)
were
administered
to
test
animals
(
7­
8
weeks
old;
10/
group)
daily
for
five
days,
off
for
two
days,
and
then
again
daily
for
5
days.
The
high
dose
was
selected
on
the
basis
that
it
had
previously
been
demonstrated
to
be
carcinogenic
in
female
mice.
Bromoform
was
administered
in
drinking
water
at
approximately
75%
of
the
saturation
level,
resulting
in
an
average
daily
dose
of
1.19
mmol/
kg
(
301
mg/
kg).
The
mice
were
sacrificed
24
hours
after
the
last
gavage
dose
and
the
livers
were
removed,
weighed,
and
processed
for
histopathological
examination,
proliferating
cell
nuclear
antigen
­
labeling
index
(
PCNA­
LI)
analysis,
and
determination
of
c­
myc
methylation
status.
For
histopathological
analysis,
stained
liver
sections
were
evaluated
for
toxicity
using
a
semiquantitative
procedure
using
the
following
severity
scoring
system:
Grade
1
consisted
of
mid
lobular
ballooning
hepatocytes;
Grade
2
consisted
of
mid
lobular
ballooning
hepatocytes
extending
to
the
central
vein;
Grade
3
consisted
of
centrilobular
necrosis
with
ballooning
hepatocytes;
and
Grade
4
consisted
of
necrosis
extending
from
the
central
vein
to
the
mid
lobule
zone.
A
significant,
dose­
dependent
increase
in
relative
liver
weight
was
observed
in
animals
dosed
by
gavage;
however,
relative
liver
weight
was
unaffected
in
animals
administered
the
compound
in
drinking
water,
when
compared
to
controls.
At
the
low
gavage
dose,
liver
toxicity
consisted
mainly
of
a
Grade
1
response.
At
the
high
dose,
liver
toxicity
consisted
mainly
of
a
Grade
2
response.
No
incidence
data
were
provided
in
the
study
report,
nor
were
severity
data
presented
for
the
control
group.
The
histopathology
findings
for
animals
receiving
bromoform
in
the
drinking
water
were
similar
to
those
observed
in
the
low
dose
gavage
group.
Bromoform
administered
by
gavage
caused
a
significant,
dose­
dependent
increase
in
the
PCNA­
LI.
Bromoform
also
significantly
enhanced
cell
proliferation
when
the
compound
was
administered
in
drinking
water.
Administration
of
bromoform
by
gavage
or
in
drinking
water
decreased
methylation
of
the
c­
myc
gene.
A
LOAEL
of
200
mg/
kg,
the
lowest
dose
tested,
was
identified
on
the
basis
of
liver
toxicity
in
gavaged
animals.
Draft
­
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or
Quote
February
20,
2003
V
­
28
C.
Subchronic
Exposure
This
section
addresses
studies
of
brominated
trihalomethanes
that
are
of
approximately
90
days
in
duration.
Table
V­
4
summarizes
the
details
of
these
subchronic
studies.
Draft
­
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or
Quote
February
20,
2003
V
­
29
Table
V­
4
Summary
of
Subchronic
Toxicity
Studies
for
Brominated
Trihalomethanes
Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Bromodichloromethane
Oral
Exposure
Chu
et
al.
(
1982b)
Rat
SD*
Drinking
water
M
20
90
days
0
0.57
6.5
53
212
Non
dose­
dependent
hepatic
and
thyroid
lesions
Chu
et
al.
(
1982b)
Rat
SD
Drinking
water
F
20
90
days
0
0.75
6.9
57
219
Non
dose­
dependent
hepatic
and
thyroid
lesions
NTP
(
1987)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
10
13
weeks
(
5
d/
wk)
0
19
38
75
(
NOAEL)
150
(
LOAEL)
300
Reduced
body
weight
gain
NTP
(
1987)
Mouse
B6C3F1
Gavage
(
corn
oil)
M
10
13
weeks
(
5
d/
wk)
0
6.3
13
25
50
(
NOAEL)
100
(
LOAEL)
Focal
necrosis
of
proximal
renal
tubular
epithelium
NTP
(
1987)
Mouse
B6C3F1
Gavage
(
corn
oil)
F
10
13
weeks
(
5
d/
wk)
0
25
50
100
(
NOAEL)
200
(
LOAEL)
400
Hepatic
microgranulomas
Inhalation
Exposure
Torti
et
al.
2001
Mouse
C57BL/
6
FVB/
N
p53
(
heterozygous
Vapor
M
Not
reported
13
weeks
(
6
h/
day)
0
ppm
0.5
ppm
3
ppm
10
ppm
15
ppm
Text
reported
minimal
cortical
scarring
and
occasional
regenerative
tubules
in
C56BL/
6
mice
and
mild
renal
cortical
tubular
karyocytomegaly.
Concentrations
at
which
these
effects
occurred
were
not
reported.
Table
V­
4
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
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or
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February
20,
2003
V
­
30
Dibromochloromethane
Chu
et
al.
(
1982b)
Rat
SD
Drinking
water
M
20
90
days
0
0.57
6.1
49
(
NOAEL)
224
(
LOAEL)
Hepatic
lesions
Chu
et
al.
(
1982b)
Rat
SD
Drinking
water
F
20
90
days
0
0.64
6.9
55
(
NOAEL)
236
(
LOAEL)
Hepatic
lesions
NTP
(
1985)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
10
13
weeks
(
5
d/
wk)
0
15
30
(
NOAEL)
60
(
LOAEL)
125
250
Hepatic
vacuolization
indicative
of
fatty
metamorphosis
(
males)

NTP
(
1985)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
10
13
weeks
(
5
d/
wk)
0
15
30
60
125
(
NOAEL)
250
(
LOAEL)
Fatty
liver
and
toxic
nephropathy
in
males
Daniel
et
al.
(
1990)
Rat
SD
Gavage
(
corn
oil)
M,
F
10
90
days
0
50
(
LOAEL)
100
200
Hepatic
vacuolization
(
males);
renal
lesions
(
females)

Bromoform
Chu
et
al.
(
1982b)
Rat
SD
Drinking
water
M
20
90
days
0
0.65
6.1
57
(
NOAEL)
218
(
LOAEL)
Hepatic
lesions
and
vacuolation
Chu
et
al.
(
1982b)
Rat
SD
Drinking
water
F
20
90
days
0
0.64
6.9
55
(
NOAEL)
283
(
LOAEL)
Hepatic
lesions
and
vacuolation
Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
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or
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February
20,
2003
V
­
31
NTP
(
1989a)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
10
13
weeks
(
5
d/
wk)
0
12
25
(
NOAEL)
50
(
LOAEL)
100
200
Hepatic
vacuolation
in
males
NTP
(
1989a)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
10
13
weeks
(
5
d/
wk)
0
25
50
100
(
NOAEL)
200
(
LOAEL)
400
Hepatic
vacuolation
in
males
*
SD,
Sprague­
Dawley
1.
Bromodichloromethane
Chu
et
al.
(
1982b)
administered
bromodichloromethane
to
male
and
female
weanling
Sprague­
Dawley
rats
(
20/
sex/
dose)
in
drinking
water
at
levels
of
0,
5,
50,
500,
or
2,500
ppm
for
90
days.
Half
of
each
group
(
10/
sex/
dose)
was
sacrificed
at
the
end
of
the
exposure
period,
and
the
remaining
animals
were
given
tap
water
for
another
90
days.
As
calculated
by
the
authors
(
using
data
on
water
consumption
and
the
average
initial
and
final
body
weights
in
the
vehicle
controls
and
the
high
dose
groups),
these
levels
corresponded
to
doses
of
approximately
0,
0.57,
6.5,
53,
and
212
mg/
kg­
day
for
males
and
0,
0.75,
6.9,
57,
and
219
mg/
kg­
day
for
females.
At
2,500
ppm,
food
consumption
was
significantly
depressed
and
significant
growth
suppression
occurred
in
both
males
and
females.
Mild
histologic
changes
were
observed
in
the
liver
and
thyroid
of
the
male
animals.
Neither
incidence
nor
severity
were
clearly
dose­
related.
Specifically,
the
incidence
of
hepatic
lesions
was
increased
in
males
at
concentrations
equal
to
or
greater
than
50
ppm,
with
similar
statistically
significant
increases
in
the
severity
of
these
lesions
in
these
dose
groups
compared
to
the
control.
The
author
noted
that
the
hepatic
lesions
were
mild
and
similar
to
the
control
following
the
90­
day
recovery
period.
Increased
incidence
of
thyroid
lesions
was
also
observed
in
males
at
concentrations
equal
to
or
greater
than
50
ppm.
The
severity
of
these
lesions
was
similar
to
that
observed
in
the
control
group.
These
lesions
were
also
mild
and
similar
in
nature
to
those
of
the
control
after
the
90­
day
recovery
period.
The
incidence
of
hepatic
lesions
in
the
female
treatment
groups
(
3­
5/
10)
was
slightly
increased
compared
to
that
of
the
control
group
(
0/
10)
with
the
severity
significantly
increased
in
the
50
and
2,500
ppm
treatment
groups,
but
not
in
the
500
ppm
group.
No
significant
numbers
of
females
were
reported
as
having
thyroid
lesions.
Lack
of
a
clear
dose­
response
relationship
for
either
incidence
or
severity
of
lesions
prevented
identification
of
reliable
NOAEL
or
LOAEL
values.

NTP
(
1987)
administered
doses
of
0,
19,
38,
75,
150,
or
300
mg/
kg­
day
of
bromodichloromethane
to
male
and
female
F344/
N
rats
(
10/
sex/
dose)
by
gavage
in
corn
oil
for
5
days/
week
for
13
weeks.
The
low­
dose
group
was
administered
1.9
mg/
kg­
day
for
the
first
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
32
3
weeks
of
the
study.
A
necropsy
was
performed
on
all
animals.
Before
study
termination,
50%
of
the
males
and
20%
of
the
females
in
the
high­
dose
group
died.
Although
food
consumption
was
not
recorded,
animals
in
the
high­
dose
groups
appeared
to
eat
less
food.
These
animals
were
also
emaciated.
At
300
mg/
kg­
day,
final
body
weights
of
the
males
and
females
were
decreased
by
55%
and
32%,
respectively,
relative
to
the
controls.
At
150
mg/
kg­
day,
final
body
weights
of
the
males
and
females
were
decreased
by
30%
and
12%,
respectively,
relative
to
the
controls.
Treatment­
related
lesions
were
observed
only
at
the
high
dose.
At
300
mg/
kg­
day
in
males,
centrilobular
degeneration
of
the
liver
and
occasional
necrotic
cells
were
observed
in
4/
9
animals.
Mild
bile
duct
hyperplasia
was
also
observed
in
these
animals.
Kidney
lesions
in
high­
dose
males
consisted
of
degeneration
of
renal
proximal
tubular
epithelial
cells
(
4/
9)
and
definite
foci
of
coagulative
necrosis
of
the
tubular
epithelium
(
2/
9).
High­
dose
males
(
4/
9)
also
exhibited
lymphoid
degeneration
of
the
thymus,
spleen,
and
lymph
nodes,
and
mild
to
moderate
atrophy
of
the
seminal
vesicles
and/
or
prostate.
Enlarged
hepatocytes
were
observed
in
females
(
2/
9)
at
300
mg/
kg­
day.
Although
degeneration
of
the
spleen,
thymus,
and
lymph
nodes
was
noted
in
high­
dose
females,
the
extent
of
the
atrophy
was
much
less
than
that
observed
in
males.
This
study
identified
a
NOAEL
of
75
mg/
kg­
day
and
a
LOAEL
of
150
mg/
kg­
day
based
on
reduced
body
weight
gain.

In
a
parallel
experiment,
NTP
(
1987)
administered
bromodichloromethane
in
corn
oil
by
gavage
to
male
and
female
B6C3F
1
mice
(
10/
sex/
dose)
for
5
days/
week
for
13
weeks.
Doses
were
0,
6.25,
12.5,
25,
50,
or
100
mg/
kg­
day
for
males
and
0,
25,
50,
100,
200,
or
400
mg/
kgday
for
females.
All
animals
survived
to
the
end
of
the
study.
The
final
body
weights
of
highdose
males
were
decreased
by
9%
relative
to
the
controls.
The
final
body
weights
of
females
that
received
200
and
400
mg/
kg­
day
were
decreased
5%
and
6%,
respectively,
relative
to
the
controls.
No
treatment­
related
clinical
signs
were
noted.
Treatment­
related
lesions
were
observed
only
at
100
mg/
kg­
day
in
males
and
at
200
and
400
mg/
kg­
day
in
females.
Kidney
lesions
in
high­
dose
males
included
focal
necrosis
of
the
proximal
renal
tubular
epithelium
(
6/
10)
and
nephrosis
of
minimal
severity
(
2/
10).
Microgranulomas
were
observed
in
the
liver
of
70%
of
the
females
that
received
the
200
mg/
kg­
day
dose.
NOAEL
and
LOAEL
values
for
female
mice
were
100
and
200
mg/
kg­
day,
respectively,
based
on
occurrence
of
microgranulomas.
This
study
identified
a
NOAEL
of
50
mg/
kg­
day
and
a
LOAEL
of
100
mg/
kg­
day
for
male
mice
on
the
basis
of
liver
histopathology.

Torti
et
al.
(
2001)
reported
results
from
a
13­
week
interim
sacrifice
conducted
as
part
of
an
inhalation
cancer
bioassay
in
p53
heterozygous
C57BL/
6
and
FVB/
N
male
mice.
Test
animals
were
exposed
to
vapor
concentrations
of
0,
0.5,
3,
10,
or
15
ppm,
6
hours/
day
for
13
weeks.
No
exposure­
related
effects
were
noted
for
mortality,
morbidity,
relative
body
weight,
relative
kidney
or
liver
weight,
or
cell
proliferation
in
liver,
kidney
or
bladder.
Histopathologic
lesions
were
limited
to
the
kidney.
The
study
authors
reported
minimal
cortical
scarring
and
occasional
regenerative
tubules
in
the
C57BL/
6
strain.
The
only
lesion
reported
for
the
FVB/
N
strain
was
limited
to
mild
renal
cortical
tubular
karyocytomegaly.
No
incidence
data
were
presented
for
these
lesions
and
the
concentrations
at
which
they
occurred
was
not
stated.

2.
Dibromochloromethane
Draft
­
Do
Not
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or
Quote
February
20,
2003
V
­
33
Chu
et
al.
(
1982b)
administered
dibromochloromethane
to
male
and
female
weanling
Sprague­
Dawley
rats
(
20/
sex/
dose)
in
drinking
water
at
levels
of
0,
5,
50,
500,
or
2,500
ppm
for
90
days.
Half
of
each
group
(
10/
sex/
dose)
was
sacrificed
at
the
end
of
the
exposure
period,
and
the
remaining
animals
were
given
tap
water
for
another
90
days.
Based
on
calculations
by
the
authors,
these
levels
corresponded
to
doses
of
approximately
0,
0.57,
6.1,
49,
and
224
mg/
kg­
day
for
males
and
0,
0.64,
6.9,
55,
and
236
mg/
kg­
day
for
females.
At
2,500
ppm,
food
consumption
was
depressed
in
both
males
and
females,
with
the
decrease
reaching
statistical
significance
in
the
males.
Body
weight
gain
was
also
decreased
at
the
high­
dose,
but
not
significantly.
Mild
histologic
changes
occurred
in
the
liver
and
thyroid
in
both
males
and
females.
Neither
the
incidence
nor
severity
exhibited
clear
dose­
response
trends
with
the
possible
exception
of
the
incidence
and
severity
of
hepatic
lesions
in
the
males.
The
severity
of
hepatic
lesions
was
significantly
increased
at
50
ppm
in
females
and
at
2,500
ppm
in
both
males
and
females.
Hepatic
lesions
included
increased
cytoplasmic
volume
and
vacuolation
due
to
fatty
infiltration.
Lesions
of
the
thyroid
included
decreased
follicular
size
and
colloid
density
and
occasional
focal
collapse
of
follicles.
The
severity
of
these
lesions
was
not
significantly
different
from
that
of
the
control.
The
authors
noted
that
histological
changes
were
mild
and
similar
to
controls
when
evaluated
after
the
90­
day
recovery
period.
These
data
identified
a
NOAEL
of
49
mg/
kg­
day
and
a
LOAEL
of
224
mg/
kg­
day
for
males,
and
a
NOAEL
of
55
mg/
kg­
day
and
a
LOAEL
of
236
mg/
kg­
day
for
females.

NTP
(
1985)
administered
dibromochloromethane
by
gavage
in
corn
oil
to
male
and
female
F344/
N
rats
(
10/
dose/
sex).
Doses
of
0,
15,
30,
60,
125,
or
250
mg/
kg
were
given
5
days/
week
for
13
weeks.
Animals
were
weighed
weekly.
All
animals
were
submitted
for
gross
necropsy,
while
histopathology
was
conducted
on
animals
in
the
control
and
high­
dose
groups
with
the
exception
that
the
liver
was
examined
in
all
males
and
in
females
at
125
mg/
kg­
day,
and
that
the
kidney
and
salivary
glands
were
examined
in
males
and
females
at
125
mg/
kg­
day.
Only
one
male
and
one
female
in
the
high­
dose
group
survived,
with
most
deaths
occurring
during
weeks
8
to
10.
At
125
mg/
kg­
day,
final
body
weights
of
males
were
decreased
7%
relative
to
controls.
Histopathological
examination
revealed
severe
lesions
and
necrosis
in
kidney,
liver,
and
salivary
glands,
primarily
at
the
high­
dose.
Males,
however,
exhibited
a
dose­
dependent
increase
in
the
frequency
of
clear
cytoplasmic
vacuoles
indicative
of
fatty
metamorphosis
in
the
liver;
this
effect
was
statistically
significant
at
doses
of
60
mg/
kg­
day
or
higher.
On
this
basis,
this
study
identified
a
NOAEL
of
30
mg/
kg­
day
and
a
LOAEL
of
60
mg/
kg­
day
in
rats
for
dibromochloromethane.
BMD
modeling
was
conducted
on
the
incidence
of
fatty
metamorphosis
in
male
rats.

NTP
(
1985)
performed
a
similar
13­
week
gavage
study
with
dibromochloromethane
in
male
and
female
B6C3F
1
mice
(
10/
sex/
dose).
The
doses
and
dosing
schedule
were
the
same
as
for
the
rat
study.
No
treatment­
related
effects
on
body
weight
or
histopathology
were
observed
at
doses
of
125
mg/
kg­
day
or
lower.
At
the
high
dose,
final
body
weights
of
males
and
females
were
decreased
by
6%
relative
to
controls.
Fatty
metamorphosis
of
the
liver
and
toxic
nephropathy
were
observed
in
high­
dose
males,
but
not
in
high­
dose
females.
This
study
identified
a
NOAEL
of
125
mg/
kg­
day
and
a
LOAEL
of
250
mg/
kg­
day
in
mice.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
34
Daniel
et
al.
(
1990)
administered
gavage
doses
(
in
corn
oil)
of
0,
50,
100,
or
200
mg/
kgday
of
dibromochloromethane
to
male
and
female
Sprague­
Dawley
rats
(
10/
sex/
dose)
for
90
consecutive
days.
Individual
dosages
were
adjusted
weekly
based
on
individual
body
weights.
During
the
final
week
of
the
study,
urinalysis
was
conducted
following
an
overnight
fast.
Ophthalmoscopic
examinations
were
performed
prior
to
treatment
and
during
the
last
week
of
the
study.
Hematology,
serum
clinical
chemistry,
and
a
thorough
histopathologic
examination
were
also
conducted.
No
deaths,
clinical
signs
of
toxicity,
or
treatment­
related
changes
in
the
ophthalmoscopic
examinations
or
hematology
were
observed.
Final
body
weights
were
significantly
reduced
in
the
high­
dose
groups
by
32%
in
males
and
by
13%
in
females.
Body
weight
decreases
in
the
other
groups
were
less
than
10%
of
control
weights.
A
dose­
related
increase
was
observed
in
liver
weight
in
females
that
reached
statistical
significance
at
the
high
dose.
Clinical
chemistry
values
indicative
of
hepatotoxicity
and
suggestive
of
nephrotoxicity
included
increased
levels
of
alkaline
phosphatase
(
high­
dose
males
and
females),
ALT
(
mid­
and
high­
dose
males),
and
creatinine
(
mid­
and
high­
dose
males
and
high­
dose
females),
and
decreased
potassium
levels
(
high­
dose
males).
Centrilobular
lipidosis
(
vacuolization)
was
observed
in
the
liver
of
almost
all
high­
dose
males
and
females
and
all
mid­
and
low­
dose
males
(
with
one
exception
at
each
level),
but
in
only
one
mid­
dose
female.
The
severity
of
the
effect
was
dose­
related.
Centrilobular
necrosis
was
also
observed
in
high­
dose
males
and
females.
Slight­
to­
moderate
degeneration
within
the
kidney
proximal
tubular
cells
occurred
in
all
high­
dose
males
and
females
and
to
a
lesser
extent
in
mid­
dose
males
and
low­
and
mid­
dose
females.
Based
on
the
liver
histopathology
in
males
and
kidney
histopathology
in
females,
the
LOAEL
for
dibromochloromethane
in
this
study
was
50
mg/
kg­
day.

3.
Bromoform
Chu
et
al.
(
1982b)
administered
bromoform
to
male
and
female
weanling
Sprague­
Dawley
rats
(
20
rats/
sex/
group)
for
90
days
in
drinking
water
at
levels
of
0,
5,
50,
500,
or
2,500
ppm.
Half
of
each
group
(
10/
sex/
dose)
was
sacrificed
at
the
end
of
the
exposure
period,
and
the
remaining
animals
were
given
tap
water
for
another
90
days.
Based
on
calculations
by
the
authors,
these
levels
corresponded
to
doses
of
approximately
0,
0.65,
6.1,
57,
and
218
mg/
kg­
day
for
males
and
0,
0.64,
6.9,
55,
and
283
mg/
kg­
day
for
females.
At
2,500
ppm,
food
consumption
was
depressed
in
both
males
and
females,
with
the
decrease
reaching
statistical
significance
in
males.
Body
weight
gain
was
also
decreased
at
the
high­
dose,
but
not
significantly.
Lymphocyte
counts
were
significantly
decreased
in
high­
dose
males
and
females
when
evaluated
90­
days
after
cessation
of
treatment.
The
only
change
in
serum
biochemistry
was
a
significant
decrease
in
LDH
in
both
males
and
females
at
the
high
dose.
This
effect
was
also
noted
90­
days
after
cessation
of
treatment.
Mild
histologic
changes
occurred
in
the
liver
and
thyroid
of
male
and
female
animals.
Although
neither
incidence
nor
severity
were
clearly
dose­
related,
these
parameters
did
tend
to
increase
with
dose.
The
severity
of
hepatic
lesions
was
significantly
increased
in
high­
dose
males
and
in
females
at
500
and
2,500
ppm.
Hepatic
lesions
included
increased
cytoplasmic
volume
and
vacuolation
due
to
fatty
infiltration.
Lesions
of
the
thyroid
included
decreased
follicular
size
and
colloid
density
and
occasional
focal
collapse
of
follicles.
The
severity
of
these
lesions
in
the
treated
animals
was
not
significantly
different
from
that
in
the
controls.
Although
the
authors
noted
that
histologic
changes
were
mild
and
similar
to
controls
when
evaluated
after
the
90­
day
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
35
recovery
period,
males
in
the
high­
dose
group
continued
to
exhibit
an
increased
incidence
of
hepatic
lesions
with
greater
severity
relative
to
the
control.
These
data
identified
a
NOAEL
of
57
mg/
kg­
day
and
a
LOAEL
of
218
mg/
kg­
day
for
males,
and
a
NOAEL
of
55
mg/
kg­
day
and
a
LOAEL
of
283
mg/
kg­
day
for
females.

NTP
(
1989a)
exposed
male
and
female
F344/
N
rats
to
bromoform
by
gavage
for
5
days/
week
for
13
weeks.
Animals
(
10/
sex/
dose)
received
doses
of
0,
12,
25,
50,
100,
or
200
mg/
kg­
day.
None
of
the
rats
died
before
the
end
of
the
study,
and
body
weights
were
not
significantly
affected.
All
high­
dose
animals,
as
well
as
males
dosed
with
100
mg/
kg­
day,
were
lethargic.
At
sacrifice,
tissues
were
examined
for
gross
and
histologic
changes.
A
dosedependent
increase
in
the
frequency
of
hepatocellular
vacuolation
was
observed
in
male
rats,
which
reached
statistical
significance
at
50
mg/
kg­
day
(
IRIS,
1993b).
These
hepatic
effects
were
not
observed
in
females.
This
study
identified
a
NOAEL
of
25
mg/
kg­
day
and
a
LOAEL
of
50
mg/
kg­
day,
on
the
basis
of
the
hepatic
vacuolation
seen
in
male
rats.

In
a
parallel
study,
NTP
(
1989a)
exposed
male
and
female
B6C3F
1
mice
to
bromoform
by
gavage
for
5
days/
week
for
13
weeks.
Animals
(
10/
sex/
dose)
received
doses
of
0,
25,
50,
100,
200,
or
400
mg/
kg­
day.
One
female
died
at
100
mg/
kg­
day,
but
no
other
deaths
at
any
other
dose
level
occurred.
At
sacrifice,
tissues
were
examined
for
gross
and
histologic
changes.
Body
weights
were
not
significantly
affected,
although
males
receiving
400
mg/
kg­
day
had
body
weights
about
8%
less
than
controls.
A
dose­
related
increase
in
the
number
of
hepatocellular
vacuoles
was
seen
in
male
mice
(
incidence
of
5/
10
at
200
mg/
kg
and
8/
10
at
400
mg/
kg
reported
in
text;
incidence
in
controls
not
explicitly
stated),
but
not
in
females.
Based
on
hepatocellular
vacuolation,
this
study
identified
a
NOAEL
of
100
mg/
kg­
day
and
a
LOAEL
of
200
mg/
kg­
day
in
male
mice.

D.
Chronic
Exposure
This
section
addresses
studies
on
the
health
effects
of
brominated
trihalomethanes
that
are
of
one
to
two
years
in
duration.
These
studies
are
summarized
in
Table
V­
5.

1.
Bromodichloromethane
Tobe
et
al.
(
1982)
evaluated
the
chronic
effects
of
bromodichloromethane
administered
in
the
diet
to
male
and
female
Slc:
Wistar
SPF
rats
(
40/
sex/
group)
for
24
months.
The
histopathology
data
for
the
animals
exposed
to
bromodichloromethane
in
this
study
were
reported
by
Aida
et
al.
(
1992b).
The
animals
were
5
weeks
old
at
the
start
of
the
study
and
weighed
approximately
100
g.
Bromodichloromethane
was
microencapsulated,
and
an
appropriate
amount
was
mixed
with
powdered
feed.
The
concentrations
administered
were
0.0.
0.014,
0.055,
or
0.22%.
Control
groups
(
70
rats/
sex)
received
microcapsules
without
the
test
compound.
Body
weight
and
food
consumption
were
monitored
weekly
for
the
first
6
months,
every
2
weeks
from
6
to
12
months,
and
every
4
weeks
during
the
second
year
of
the
study.
Interim
sacrifices
of
at
least
9
animals/
sex/
control
group
and
5
animals/
sex/
dose
group
were
conducted
at
6,
12,
and
18
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
36
months.
All
surviving
animals
were
sacrificed
at
2
years.
At
each
time
of
sacrifice,
necropsies,
hematology,
and
serum
biochemistry
were
conducted.
Based
on
mean
food
intakes,
the
reported
average
doses
were
approximately
0,
6,
26,
or
138
mg/
kg­
day
for
males
and
0,
8,
32,
or
168
mg/
kg­
day
for
females
(
Aida
et
al.,
1992b).
Marked
suppression
of
body
weight
gain
was
seen
in
males
and
females
of
the
high­
dose
group.
Males
and
females
of
the
high­
dose
group
exhibited
mild
piloerection
and
emaciation.
Relative
liver
weight
was
significantly
increased
in
the
mid­
and
high­
dose
groups,
while
relative
kidney
weight
was
significantly
increased
only
in
the
high­
dose
groups.
At
18
months,
dose­
dependent
reductions
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
37
Table
V­
5
Summary
of
Chronic
Toxicity
Studies
for
Brominated
Trihalomethanes
Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Bromodichloromethane
NTP
(
1987)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
50
2
years
(
5
d/
wk)
0
50
(
LOAEL)
100
Renal
and
hepatic
histopathology
NTP
(
1987)
Mouse
B6C3F1
Gavage
(
corn
oil)
M
50
2
years
(
5
d/
wk)
0
25
(
LOAEL)
50
Renal
and
hepatic
histopathology
NTP
(
1987)
Mouse
B6C3F1
Gavage
(
corn
oil)
F
50
2
years
(
5
d/
wk)
0
75
(
LOAEL)
150
Reduced
body
weight
gain
Aida
et
al.
(
1992b);
Tobe
et
al.
(
1982)
Rat
Wistar
Diet
M
40
2
years
0
6
(
LOAEL)
26
138
Hepatic
vacuolization,
serum
chemistry
Aida
et
al.
(
1992b);
Tobe
et
al.
(
1982)
Rat
Wistar
Diet
F
40
2
years
0
8
(
NOAEL)
32
(
LOAEL)
168
Hepatic
vacuolization,
serum
chemistry
Klinefelter
et
al.
(
1995)
Rat
F344
Drinking
water
M
7
1
year
0
22
39
(
NOAEL)
No
evidence
of
treatment­
related
histopathological
or
organ
weight
effects
(
see
reproductive
effects,
section
V.
E
for
additional
data
from
this
study)

Dibromochloromethane
Tobe
et
al.
(
1982)
Rat
Wistar
SPF
Diet
M
40
2
years
0
12
(
NOAEL)
49
(
LOAEL)
196
Serum
biochemistry,
liver
appearance
at
necropsy;
decreased
body
weight
gain
Tobe
et
al.
(
1982)
Rat
Wistar
SPF
Diet
F
40
2
years
0
17
(
NOAEL)
70
(
LOAEL)
278
Serum
biochemistry,
liver
appearance
at
necropsy;
decreased
body
weight
gain
NTP
(
1985)
Rat
F344
Gavage
(
corn
oil)
M,
F
50
2
years
(
5
d/
wk)
0
40
(
LOAEL)
80
Histologic
changes
in
liver,
including
fat
accumulation
and
ground
glass
appearance,
and
altered
basophilic
staining
NTP
(
1985)
Mouse
B6C3F1
Gavage
(
corn
oil)
M,
F
50
105
weeks
(
5
d/
wk)
0
50
(
LOAEL)
100
Fatty
metamorphosis
in
liver
and
follicular
cell
hyperplasia
in
thyroid
Table
V­
5
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
38
Bromoform
Tobe
et
al.
(
1982)
Rat
Wistar
SPF
Diet
M
40
2
years
0
22
(
NOAEL)
90
(
LOAEL)
364
Enzyme
changes
and
altered
liver
appearance
at
necropsy
Tobe
et
al.
(
1982)
Rat
Wistar
SPF
Diet
F
40
2
years
0
38
(
NOAEL)
152
(
LOAEL)
619
Enzyme
changes
and
altered
liver
appearance
at
necropsy
NTP
(
1989a)
Rat
F344/
N
Gavage
(
corn
oil)
M,
F
50
103
weeks
(
5
day/
wk)
0
100
(
LOAEL)
200
Decreased
body
weight,
lethargy,
mild
hepatotoxicity
NTP
(
1989a)
Mouse
B6C3F1
Gavage
(
corn
oil)
M
50
103
weeks
(
5
day/
wk)
0
50
100
(
NOAEL)
No
observed
effects
on
body
weight
or
hepatotoxicity
NTP
(
1989a)
Mouse
B6C3F1
Gavage
(
corn
oil)
F
50
103
weeks
(
5
day/
wk)
0
100
(
LOAEL)
200
Decreased
body
weight,
minimal
to
mild
fatty
changes
in
liver
*
SD,
Sprague­
Dawley
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
39
in
serum
cholinesterase
activity
and
increases
in
 ­
glutamyl
transpeptidase
(
 ­
GTP)
activity
(
indicative
of
bile
duct
proliferation)
were
observed
in
males,
with
the
changes
significant
at
the
high
dose.
The
mid­
and
high­
dose
males
also
displayed
a
27%
and
65%
reduction,
respectively,
in
total
serum
triglycerides
(
T­
Gly)
levels
when
compared
to
the
control
group.
At
18
months,
serum
cholinesterase
levels
were
significantly
decreased,
while
total
cholesterol
levels
were
significantly
increased,
in
all
dose
groups
for
the
treated
females.
Serum
T­
Gly
levels
(
decreased)
and
 ­
GTP
activity
(
increased)
were
also
reported
to
deviate
significantly
from
control
values
in
the
mid­
and
high­
dose
females.
The
most
sensitive
markers
at
24
months
were
T­
GLY
and
serum
cholinesterase,
with
significant
changes
seen
in
all
of
the
male
treatment
groups.
Gross
necropsy
revealed
dose­
related
yellowing
and
roughening
of
the
liver
surface.
Treatment­
related
lesions
were
limited
to
the
liver.
At
24
months,
fatty
degeneration
and
granuloma
were
observed
in
all
dose
groups
with
the
exception
of
granulomas
in
low­
dose
females.
Specifically,
fatty
degeneration
and
granulomas
were
observed
in
low­
dose
males,
but
not
control
males,
and
fatty
degeneration
was
observed
in
low­
dose
females
at
a
higher
rate
(
8/
19)
than
in
control
females
(
2/
32).
Cholangiofibrosis
was
also
observed
in
the
high­
dose
groups.
Bile
duct
proliferation
was
observed
in
most
high­
dose
animals
at
6
months,
and
was
prevalent
in
the
controls
and
all
dose
groups
by
24
months.
Histopathology
was
observed
in
all
dose
groups
as
early
as
6
months
with
the
exception
of
low­
dose
females.
Based
on
the
results
of
Tobe
et
al.
(
1982)
alone,
the
NOAEL
was
6
mg/
kg­
day
in
males
and
8
mg/
kg­
day
in
females.
The
LOAEL
was
identified
as
26
(
males)
to
32
(
females)
mg/
kg­
day,
based
on
serum
enzyme
changes
and
altered
liver
appearance.
Based
on
the
histopathology
data
reported
for
this
study
by
Aida
et
al.
(
1992b),
however,
the
entire
study
identified
a
LOAEL
of
6
mg/
kg­
day
in
male
rats
and
8
mg/
kg­
day
in
female
rats.

NTP
(
1987)
administered
doses
of
0,
50,
or
100
mg/
kg­
day
of
bromodichloromethane
in
corn
oil
by
gavage
to
male
and
female
F344/
N
rats
(
50/
sex/
dose),
5
days/
week
for
102
weeks.
The
authors
observed
all
animals
for
clinical
signs
and
recorded
body
weights
(
by
cage)
once
per
week
for
the
first
12
weeks
of
the
study
and
once
per
month
thereafter.
A
necropsy
was
performed
on
all
animals,
including
those
found
dead,
unless
they
were
excessively
autolyzed
or
cannibalized.
During
necropsy,
all
organs
and
tissues
were
examined
for
grossly
visible
lesions.
Complete
histopathology
was
performed
on
all
female
rats
and
on
high­
dose
and
vehicle­
control
male
rats.
Male
rats
in
the
low­
dose
group
that
died
early
in
the
study
were
also
examined
histologically.
Survival
of
dosed
rats
was
comparable
to
that
of
vehicle
controls.
Mean
body
weight
of
high­
dose
male
and
female
rats
was
decreased
during
the
last
1.5
years
of
the
study;
body
weight
gain
of
high­
dose
male
and
female
rats
was
86%
and
70%
of
the
corresponding
vehicle­
control
values.
Body
weight
gain
of
low­
dose
male
and
female
rats
was
comparable
to
that
of
the
vehicle­
control
group.
No
treatment­
related
clinical
signs
were
observed.
In
males,
treatment­
related
nonneoplastic
effects
included
renal
cytomegaly,
tubular
cell
hyperplasia,
hepatic
necrosis,
and
fatty
metamorphosis.
In
females,
changes
included
eosinophilic
cytoplasmic
change,
clear
cell
change,
focal
cellular
change,
fatty
metamorphosis
of
the
liver,
and
tubular
cell
hyperplasia
of
the
kidney.
Based
on
these
histologic
findings,
this
study
identified
a
LOAEL
of
50
mg/
kg­
day
in
rats.

NTP
(
1987)
administered
bromodichloromethane
in
corn
oil
by
gavage
to
male
and
female
B6C3F
1
mice
(
50/
sex/
dose),
5
days/
week
for
102
weeks.
For
males,
doses
were
0,
25,
or
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
40
50
mg/
kg­
day.
For
females,
doses
were
0,
75,
or
150
mg/
kg­
day.
Final
survival
of
treated
male
mice
was
comparable
to
that
of
vehicle
controls.
At
week
84,
survival
of
female
mice
was
greater
than
50%
in
all
dose
groups.
After
week
84,
survival
of
dosed
and
vehicle­
control
female
mice
was
reduced
(
final
survival:
26/
50,
13/
50,
15/
50
for
the
0,
5,
and
50
mg/
kg­
day
groups,
respectively),
and
this
decreased
survival
was
associated
with
ovarian
abscesses
(
8/
50,
19/
47,
18/
49).
Body
weight
gain
of
high­
dose
male
mice
was
87%
of
that
of
the
vehicle­
control
group;
the
body
weight
gain
of
low­
dose
male
mice
was
comparable
to
that
of
the
vehicle­
control
group.
Mean
body
weight
of
high­
dose
female
mice
was
decreased
during
the
last
1.5
years
of
the
study.
The
body
weight
gain
was
reduced
55%
compared
to
the
controls
at
the
high
dose
and
by
25%
among
low­
dose
females.
In
males,
treatment­
related
nonneoplastic
changes
included
fatty
metamorphosis
of
the
liver,
renal
cytomegaly,
and
follicular
cell
hyperplasia
of
the
thyroid
gland.
In
females,
hyperplasia
of
the
thyroid
gland
was
observed.
This
study
identified
a
LOAEL
of
25
mg/
kg­
day,
based
on
histopathological
findings
in
male
mice.

Klinefelter
et
al.
(
1995)
reported
interim
(
52­
week)
necropsy
data
from
a
cancer
bioassay
in
which
male
F344
rats
were
administered
average
concentrations
of
0,
330,
or
620
mg/
L
bromodichloromethane
in
drinking
water.
Corresponding
doses
of
0,
22,
and
39
mg/
kg­
day
were
calculated
by
the
authors
using
water
consumption
and
body
weight
data.
For
the
interim
sacrifice,
7
animals
per
dose
group
were
killed,
and
the
testis,
epididymis,
liver,
spleen,
kidney,
thyroid,
stomach,
intestine,
and
bladder
were
evaluated
histopathologically.
Bromodichloromethane
had
no
effect
on
body
weight
or
on
the
kidney,
liver,
spleen,
or
thyroid
weight.
There
was
no
histopathological
evidence
of
bromodichloromethane­
related
noncancer
or
cancer
effects
on
any
of
the
examined
organs.
High
levels
of
nephropathy
and
interstitial
cell
hyperplasia
were
observed,
but
these
lesions
were
not
treatment­
related.
The
NOAEL
and
LOAEL
for
this
study
are
based
on
reproductive
endpoints.
These
reproductive
effects
are
summarized
in
Section
V.
E.
1.

2.
Dibromochloromethane
Tobe
et
al.
(
1982)
evaluated
the
chronic
effects
of
dibromochloromethane
administered
in
the
diet
to
male
and
female
Slc:
Wistar
SPF
rats
(
40/
sex/
group)
for
24
months.
The
animals
were
5
weeks
old
at
the
start
of
the
test
and
weighed
approximately
100
g.
Dibromochloromethane
was
microencapsulated,
and
an
appropriate
amount
was
mixed
with
powdered
feed.
Control
groups
(
70
rats/
sex)
received
microcapsules
without
test
compound.
Body
weight
and
food
consumption
were
monitored
weekly
for
the
first
6
months,
every
2
weeks
from
6
to
12
months,
and
every
4
weeks
during
the
second
year
of
the
study.
Data
were
reported
from
the
sacrifices
of
9
animals/
sex/
control
group
and
5/
sex/
dose
group
at
18
months;
all
surviving
animals
were
sacrificed
at
24
months.
Necropsies,
hematology,
and
serum
biochemistry
were
conducted
at
the
time
of
sacrifice.
No
histopathology
data
for
dibromochloromethane
have
been
published
from
this
study.
Dibromochloromethane
was
administered
at
dietary
levels
of
0.0%,
0.022%,
0.088%,
or
0.35%.
Based
on
reported
body
weights
(
150
to
475
g
for
males
and
100
to
215
g
for
females)
and
food
consumption
(
15
to
20
g/
day
for
males
and
10
to
15
g/
day
for
females),
these
levels
corresponded
to
doses
of
approximately
0,
12,
49,
and
196
mg/
kg­
day
for
males
and
0,
17,
70,
and
278
mg/
kg­
day
for
females.
Marked
suppression
of
body
weight
gain
was
seen
in
males
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V
­
41
and
females
at
the
high
dose,
and
mild
suppression
of
body
weight
gain
(
about
10%)
was
seen
in
males
and
females
at
the
mid
dose.
Decreased
T­
GLY
and
serum
cholinesterase
activity
and
increased
 ­
GTP
were
seen
in
the
mid­
and
high­
dose
males
and
females.
Yellowing
of
the
liver
surface
was
noted
in
the
mid­
and
high­
dose
groups,
and
roughening
of
the
liver
surface
was
noted
in
high­
dose
males.
This
study
suggests
NOAELs
of
12
mg/
kg­
day
(
males)
and
17
mg/
kgday
(
females),
and
LOAELs
of
49
mg/
kg­
day
(
males)
and
70
mg/
kg­
day
(
females),
based
on
serum
biochemistry
data,
decreased
body
weight,
and
gross
necropsy
findings.

NTP
(
1985)
investigated
the
chronic
oral
toxicity
of
dibromochloromethane
in
male
and
female
F344/
N
rats.
Groups
of
50
animals/
sex/
dose
were
administered
doses
of
0,
40,
or
80
mg/
kg­
day
by
gavage
in
corn
oil
for
5
days/
week
for
104
weeks.
Survival
was
comparable
in
all
dose
groups.
Body
weight
gain
was
decreased
in
high­
dose
males
after
week
20;
final
weight
gain
was
88%
of
the
control
value.
Females
in
both
dose
groups
gained
more
weight
than
did
the
controls.
Histologic
lesions
in
the
liver
were
observed
in
both
males
and
females
at
both
dose
levels.
Changes
included
fat
accumulation,
"
ground
glass"
appearance
of
the
cytoplasm,
and
altered
basophilic
staining.
This
study
identified
a
LOAEL
of
40
mg/
kg­
day
for
dibromochloromethane
in
rats.

NTP
(
1985)
performed
a
similar
chronic
oral
exposure
study
of
dibromochloromethane
toxicity
in
male
and
female
B6C3F
1
mice.
Groups
of
50
animals/
sex/
dose
were
administered
doses
of
0,
50,
or
100
mg/
kg­
day
by
gavage
in
corn
oil
for
5
days/
week
for
105
weeks.
Survival
in
females
was
not
different
from
controls,
while
survival
in
high­
dose
males
was
significantly
decreased.
An
overdosing
accident
at
week
58
killed
35/
50
male
mice
in
the
low­
dose
group,
and
this
group
was
not
considered
further.
Mean
body
weight
was
decreased
in
high­
dose
males
and
females,
but
not
in
low­
dose
females.
Treatment­
related
hepatocytomegaly
and
focal
necrosis
were
observed
in
livers
of
high­
dose
males.
Females
showed
liver
calcification
at
the
high
dose
and
fatty
metamorphosis
at
both
the
low
and
high
doses.
An
increased
incidence
of
follicular
cell
hyperplasia
in
the
thyroid
was
observed
in
low­
and
high­
dose
females
relative
to
the
control.
Thyroid
lesions
were
not
observed
in
treated
males.
This
study
identified
a
LOAEL
of
50
mg/
kgday
for
dibromochloromethane
in
mice.

3.
Bromoform
Tobe
et
al.
(
1982)
evaluated
the
chronic
effects
of
bromoform
administered
in
the
diet
to
male
and
female
Slc:
Wistar
SPF
rats
(
40/
sex/
group)
for
24
months.
The
animals
were
5
weeks
old
at
the
start
of
the
test
and
weighed
approximately
100
g.
Bromoform
was
microencapsulated,
and
administered
at
dietary
levels
of
0.0%,
0.04%,
0.16%,
or
0.65%.
Control
groups
(
70
rats/
sex)
received
microcapsules
without
test
article.
Body
weights
and
food
consumption
were
monitored
weekly
for
the
first
6
months,
every
2
weeks
from
6
to
12
months,
and
every
4
weeks
during
the
second
year
of
the
study.
Data
were
reported
from
the
sacrifices
of
9
animals/
sex
in
the
control
group
and
5/
sex/
dose
in
the
exposure
groups
at
18
months;
all
surviving
animals
were
sacrificed
at
24
months.
At
each
time
of
sacrifice,
necropsies,
hematology,
and
serum
biochemistry
were
conducted.
No
histopathology
data
for
bromoform
have
been
published
from
this
study.
Based
on
reported
body
weights
(
150
to
475
g
for
males
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V
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42
and
100
to
215
g
for
females)
and
food
consumption
(
15
to
20
g/
day
for
males
and
10
to
20
g/
day
for
females),
these
levels
corresponded
to
doses
of
about
0,
22,
90,
and
364
mg/
kg­
day
for
males
and
0,
38,
152,
and
619
mg/
kg­
day
for
females.
Marked
suppression
of
body
weight
gain
was
seen
in
males
and
females
at
the
high
dose,
and
mild
suppression
of
body
weight
gain
(
about
15%)
was
seen
in
males
and
females
at
the
mid
dose.
Dose­
related
decreases
in
non­
esterified
fatty
acids
were
observed
in
all
treated
males
and
in
females
at
the
mid
and
high
dose.
Females
also
exhibited
a
dose­
related
increase
in
levels
of
 ­
GTP
with
the
increases
significant
at
the
mid
and
high
dose.
Other
serum
biochemistry
changes
in
the
high­
dose
groups
included
decreased
serum
triglyceride
(
T­
GLY)
and
increased
AST
and
ALT
activity.
Specifically,
T­
GLY
levels
significantly
decreased
by
86%
and
80%
in
the
male
and
female
high­
dose
groups,
respectively,
by
study
termination.
AST
and
ALT
activities
at
study
termination
were
significantly
increased
1.6
to
2.6­
fold
in
animals
at
the
high
dose
compared
to
controls,
with
the
exception
that
the
increase
in
AST
activity
in
males
was
statistically
nonsignificant.
Yellowing
and
small
white
spots,
and
roughening
of
the
surface
were
seen
in
the
livers
of
the
mid­
and
high­
dose
animals.
Roughening
of
the
liver
surface
was
observed
in
the
high­
dose
groups.
Based
on
the
necropsy
findings
and
the
serum
biochemistry
data,
this
study
indicated
NOAELs
of
22
mg/
kg­
day
for
males
and
38
mg/
kgday
for
females,
and
LOAELs
of
90
mg/
kg­
day
for
males
and
152
mg/
kg­
day
for
females.

NTP
(
1989a)
exposed
male
and
female
F344/
N
rats
(
50/
sex/
group)
to
bromoform
by
gavage
in
oil
for
103
weeks
(
5
days/
week)
at
doses
of
0,
100,
or
200
mg/
kg­
day.
Animals
were
observed
for
clinical
signs
throughout
the
study
(
2
days/
week).
At
termination,
necropsy
and
histopathological
examination
were
performed
on
all
animals.
Body
weight
gain
was
decreased
by
37%
in
high­
dose
females
and
by
29%
in
high­
dose
males
relative
to
the
respective
controls.
Survival
of
the
high­
dose
males
was
also
decreased.
Both
males
and
females
were
lethargic.
Hepatic
fatty
change
and
chronic
inflammation
were
noted
in
both
males
and
females
at
both
doses,
and
minimal
necrosis
was
increased
in
high­
dose
males.
Nonneoplastic
changes
were
not
reported
in
other
tissues.
This
study
identified
a
LOAEL
of
100
mg/
kg­
day
in
both
male
and
female
rats.

NTP
(
1989a)
exposed
groups
of
50
male
B6C3F
1
mice
by
gavage
in
oil
to
doses
of
0,
50,
or
100
mg/
kg­
day
of
bromoform
for
103
weeks
(
5
days/
week).
Groups
of
50
female
mice
were
administered
doses
of
0,
100,
or
200
mg/
kg­
day.
Animals
were
observed
for
clinical
signs
2
days/
week
throughout
the
study.
At
termination,
all
animals
were
necropsied,
and
a
thorough
histological
examination
of
tissues
was
performed.
Decreased
survival
was
observed
in
females,
but
not
males.
This
was
at
least
partly
due
to
a
utero­
ovarian
infection.
No
clinical
signs
were
noted.
Body
weight
gains
were
82%
and
72%
of
the
control
values
for
low­
and
high­
dose
females,
respectively,
but
body
weight
gain
was
not
affected
in
males.
Increased
incidences
of
minimal
to
mild
fatty
changes
were
noted
in
the
livers
of
both
low­
and
high­
dose
females,
but
not
males.
Nonneoplastic
changes
were
not
detected
in
other
tissues.
This
study
identified
a
LOAEL
of
100
mg/
kg­
day
for
female
mice,
based
on
decreased
body
weight
and
fatty
changes
of
the
liver.
No
NOAEL
for
females
was
identified.
For
males,
a
NOAEL
of
100
mg/
kg­
day
was
identified.
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February
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2003
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­
43
E.
Reproductive
and
Developmental
Effects
Studies
that
have
examined
the
reproductive
and
developmental
toxicity
of
the
brominated
trihalomethanes
are
summarized
in
Table
V­
9
at
the
end
of
this
section.

1.
Bromodichloromethane
a.
Studies
in
Rats
Ruddick
et
al.
(
1983)
investigated
the
teratogenicity
and
developmental
toxicity
of
bromodichloromethane
in
Sprague­
Dawley
rats.
Pregnant
dams
(
15/
dose
group)
were
administered
0,
50,
100,
or
200
mg/
kg­
day
by
gavage
in
corn
oil
on
gestation
days
(
GD)
6
to
15.
Body
weights
were
measured
on
GD
1,
on
GD
1
through
GD
15,
and
before
and
after
fetuses
were
removed
by
caesarean
section
on
GD
22.
On
GD
22,
females
were
sacrificed
and
body
tissues
(
including
the
uterus)
were
removed
for
pathological
examination.
Females
were
evaluated
for
the
number
of
resorption
sites,
and
number
of
fetuses.
Maternal
blood
samples
were
collected
and
evaluated
for
standard
hematology
and
clinical
chemistry
parameters.
The
liver,
heart,
brain,
spleen,
and
one
kidney
were
weighed.
Standard
histopathology
was
conducted
on
control
and
high
dose
females
(
5/
group).
All
fetuses
were
individually
weighed,
and
evaluated
for
viability
and
external
malformations.
Histopathologic
examination
was
performed
on
two
pups
per
litter.
Of
the
remaining
live
fetuses,
approximately
two­
thirds
were
examined
for
skeletal
alterations
and
one­
third
for
visceral
abnormalities.

Although
15
inseminated
females
per
dose
group
were
exposed
to
bromodichloromethane
not
all
females
became
pregnant
and/
or
delivered
litters.
Therefore,
the
number
of
litters
per
dose
group
ranged
from
9
to
14.
One
animal
died
in
the
control
group,
but
no
deaths
occurred
in
any
of
the
exposed
groups.
In
the
high­
dose
group,
maternal
weight
gain
was
significantly
depressed
by
38%
as
compared
with
controls.
Although
maternal
weight
gains
were
also
reduced
in
the
low­
and
mid­
dose
groups
(
13%
and
15%,
respectively,
as
compared
with
controls),
these
differences
were
not
reported
as
statistically
significant.
Relative
maternal
liver
weight
was
significantly
increased
in
all
exposed
groups
(
110%,
110%,
and
117%
for
the
low­,
mid­,
and
high­
dose
groups,
respectively
as
compared
with
control
values).
Relative
kidney
and
brain
weights
were
also
statistically
increased
in
the
high­
dose
group
only.
These
increases
in
relative
organ
weights
may
have
been
associated
with
the
decreased
body
weight
gains
in
treated
females.
No
treatment­
related
changes
in
hematology,
clinical
chemistry,
histopathology,
number
of
resorptions,
and
the
number
of
fetuses
per
litter
were
noted.
No
differences
between
treated
and
control
groups
were
reported
for
fetal
weights,
gross
malformations
(
terata),
and
visceral
abnormalities.
However,
an
increase
in
the
incidence
of
sternebral
anomalies
was
observed
in
all
treated
groups.
The
number
of
affected
fetuses/
number
of
affected
litters
were
2/
2,
8/
4,
9/
7,
10/
6
for
the
control,
low­,
mid­,
and
high­
dose
groups,
respectively.
Statistical
significance
of
fetotoxic
endpoints
was
not
reported
by
the
study
authors.
An
independent
statistical
analysis
(
using
the
Fisher
Exact
test)
was
conducted
on
the
published
data
for
development
of
this
Criteria
Document
and
demonstrated
that
none
of
these
increases
differed
significantly
from
control
values
(
p>
0.05).
A
trend
test
showed
a
statistically
significant
dose­
related
trend
(
p=
0.03);
stepwise
Draft
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February
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2003
V
­
44
analysis
indicated
that
the
trend
became
nonsignificant
if
the
high­
dose
(
200
mg/
kg­
day)
was
omitted
from
the
analysis.
These
findings
suggest
that
the
LOAEL
and
NOAEL
for
developmental
toxicity
are
200
and
100
mg/
kg­
day,
respectively.
However,
it
should
be
noted
that
the
small
sample
sizes
(
the
sampling
unit
is
the
litter)
limited
the
statistical
power
of
the
experiment
to
detect
possible
significant
differences
at
lower
doses.
Based
on
significantly
decreased
maternal
body
weight
gain,
the
LOAEL
and
NOAEL
for
maternal
toxicity
are
200
and
100
mg/
kg­
day,
respectively.

Klinefelter
et
al.
(
1995)
evaluated
the
effects
of
bromodichloromethane
exposure
on
male
reproduction
during
a
chronic
cancer
bioassay
study
in
which
F344
rats
were
administered
bromodichloromethane
in
drinking
water
at
concentrations
of
0,
330,
or
620
mg/
L.
The
authors
estimated
the
doses
to
be
0,
22,
and
39
mg/
kg­
day.
At
52
weeks,
the
authors
conducted
an
interim
sacrifice,
which
included
an
evaluation
of
epididymal
sperm
motion
parameters
and
histopathology
of
the
testes
and
epididymides.
No
histologic
alterations
were
observed
in
any
reproductive
tissue.
Sperm
velocities
(
mean
straight­
line,
average
path,
and
curvilinear),
however,
were
significantly
decreased
at
39
mg/
kg­
day.
No
effect
on
sperm
motility
was
observed
at
22
mg/
kg­
day.
The
NOAEL
and
LOAEL
for
reproductive
effects
are
thus
22
and
39
mg/
kg­
day,
respectively.

Narotsky
et
al.
(
1997)
examined
both
the
developmental
toxicity
and
the
effect
of
dosing
vehicle
on
the
developmental
toxicity
of
bromodichloromethane.
F344
rats
(
12
to
14/
group)
were
administered
bromodichloromethane
by
gavage,
in
either
corn
oil
or
an
aqueous
vehicle
containing
10%
Emulphor
®
,
at
dose
levels
of
0,
25,
50,
or
75
mg/
kg­
day
on
GD
6
to
15.
Dams
were
allowed
to
deliver
naturally,
and
pups
were
evaluated
postnatally.
Maternal
body
weights
were
assessed
on
GD
5,
6,
8,
10,
13,
and
20,
and
all
rats
were
observed
for
clinical
signs
of
toxicity
throughout
the
test
period.
Postnatal
day
(
PND)
1
was
defined
as
GD
22
irrespective
of
the
actual
time
of
parturition.
All
pups
were
examined
externally
for
gross
malformations
and
weighed
on
PND
1
and
6.
Skeletal
and
visceral
anomalies
in
the
pups
were
not
evaluated.
Following
PND
6
examination,
the
dams
were
sacrificed
and
the
number
of
uterine
implantation
sites
per
female
was
recorded.
The
uteri
of
females
that
did
not
deliver
litters
were
stained
and
evaluated
histopathologically
to
detect
any
cases
of
full­
litter
resorption
(
FLR).
In
order
to
compare
the
kinetics
of
dosing
vehicles,
a
separate
experiment
was
conducted
in
which
pregnant
females
(
3
to
4
animals
per
vehicle
per
time
point)
were
administered
a
single
dose
of
75
mg/
kg
on
GD
6
and
whole
blood
samples
were
collected
at
30
minutes,
90
minutes,
4.5
hours,
or
24
hours
postdosing.
Following
blood
collection,
the
animals
were
sacrificed,
blood
concentrations
of
bromodichloromethane
were
measured,
and
pregnancy
status
was
confirmed
at
necropsy.

In
the
developmental
toxicity
study,
one
animal
that
received
75
mg/
kg­
day
in
corn
oil
died
before
study
termination.
In
the
mid­
and
high­
dose
groups,
clinical
signs
of
toxicity
were
evident
among
animals
administered
bromodichloromethane
in
either
dosing
vehicle.
At
75
mg/
kg­
day,
kyphosis
(
humpback)
was
observed
in
animals
receiving
the
oil
vehicle,
and
piloerection
was
observed
in
animals
receiving
either
vehicle.
At
50
mg/
kg­
day,
piloerection
was
observed
in
animals
receiving
the
aqueous
gavage,
and
chromodacryorrhea/
lacrimation
was
observed
in
animals
receiving
the
oil
gavage.
Maternal
weight
gain
was
significantly
decreased
in
Draft
­
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February
20,
2003
V
­
45
all
dosed
groups
receiving
the
aqueous
vehicle
and
in
the
50
and
75
mg/
kg­
day
groups
in
animals
receiving
the
oil
vehicle
on
GD
6
to
8
(
data
not
reported
for
other
time
periods).
Although
maternal
weight
gain
was
also
reduced
at
25
mg/
kg­
day
in
animals
given
the
oil
vehicle,
this
decrease
was
not
statistically
significant.
However,
a
two­
way
analysis
of
variance
(
ANOVA)
indicated
that
there
was
no
interaction
between
vehicle
and
dose
for
this
maternal
endpoint.
All
control
and
25
mg/
kg­
day
litters
survived
the
test
period;
however,
FLR
was
observed
at
50
and
75
mg/
kg­
day
with
both
dosing
vehicles.
Statistical
analysis
(
ANOVA)
of
FLR
incidence
showed
a
significant
vehicle­
dose
interaction.
For
females
receiving
bromodichloromethane
in
corn
oil,
FLR
was
reported
in
8
and
83%
of
the
litters
at
50
and
75
mg/
kg­
day,
respectively;
an
additional
high­
dose
litter
was
carried
to
term
but
was
delivered
late
(
GD
23),
and
all
pups
died
by
PND
6.
For
females
receiving
the
aqueous
vehicle,
FLR
was
observed
in
17
and
21%
of
the
litters
at
50
and
75
mg/
kg­
day,
respectively.
There
were
no
effects
on
gestation
length,
pre­
or
postnatal
survival,
or
pup
morphology
in
surviving
litters,
with
the
exception
noted
above
in
the
75
mg/
kgday
oil
vehicle
group.
Based
on
full
litter
resorption,
the
LOAEL
for
developmental
toxicity
is
50
mg/
kg­
day
for
both
vehicles,
and
the
corresponding
NOAEL
is
25
mg/
kg­
day.
Based
on
significantly
reduced
body
weight
gain
during
GD
6
to
8
in
dams
receiving
the
aqueous
vehicle,
the
LOAEL
for
maternal
toxicity
is
the
lowest
dose
tested,
25
mg/
kg­
day,
and
a
NOAEL
could
not
be
determined.

Analysis
of
bromodichloromethane
concentrations
in
blood
indicated
that
circulating
levels
decreased
over
time
with
both
vehicles,
but
tended
to
be
higher
following
corn
oil
administration.
Bromodichloromethane
blood
concentrations
were
thus
vehicle­
dependent
and
differed
statistically
at
both
4.5
and
24
hours
postdosing
(
mean
of
3.1
ng/
mL
versus
0.4
ng/
mL
for
oil
and
aqueous
vehicles,
respectively,
at
24
hours).
The
elimination
half­
life
of
bromodichloromethane
was
estimated
to
be
3.6
hours
when
administered
in
corn
oil
and
2.7
hours
when
given
in
the
aqueous
vehicle.

Narotsky
et
al.
(
1997)
also
calculated
both
an
ED
05
(
i.
e.,
the
effective
dose
producing
a
5%
increase
in
response
rate
above
background)
and
a
benchmark
dose
(
BMD;
as
defined
by
the
authors,
the
BMD
is
the
lower
confidence
interval
of
the
ED
05)
for
each
vehicle.
For
the
corn
oil
vehicle,
the
ED
05
and
BMD
were
48.4
and
39.3
mg/
kg­
day,
respectively.
For
the
aqueous
vehicle,
the
ED
05
and
BMD
were
33.3
and
11.3
mg/
kg­
day,
respectively.
The
study
authors
noted
that
the
greater
BMD
value
for
the
corn
oil
vehicle
seemed
counterintuitive
in
view
of
the
higher
FLR
response
rate
in
the
75
mg/
kg­
day
aqueous
vehicle
group
(
83%
for
aqueous
vehicle
versus
21%
for
corn
oil
vehicle).
However,
the
dose
response
for
bromodichloromethane­
induced
FLR
differed
markedly
between
vehicles,
and
the
response
rate
in
the
50
mg/
kg­
day
corn
oil
vehicle
group
(
8%)
closely
approximated
5%,
the
effect
level
defined
by
the
ED
05.
According
to
the
study
authors,
this
resulted
in
a
smaller
confidence
interval
around
the
ED
05
for
the
corn
oil
vehicle,
yielding
a
less
conservative
(
i.
e.,
higher)
BMD.
These
findings
are
consistent
with
the
pharmacokinetic
data
demonstrating
a
slower
elimination
of
bromodichloromethane
following
a
single
dose
of
75
mg/
kg
in
corn
oil
as
compared
with
the
same
dose
in
aqueous
vehicle,
and
suggest
that
the
influence
of
vehicle
on
FLR
rate
is
dose­
dependent.
Draft
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February
20,
2003
V
­
46
NTP
(
1998)
conducted
a
short­
term
reproductive
and
developmental
toxicity
screen
in
Sprague­
Dawley
rats
to
evaluate
the
effects
of
bromodichloromethane
administered
in
drinking
water.
The
study
was
designed
to
identify
the
physiological
endpoints
most
sensitive
to
bromodichloromethane
exposure,
and
assessed
development,
female
reproduction,
male
reproduction,
hematology,
clinical
chemistry,
and
pathology.
In
males,
the
reproductive
endpoints
evaluated
included
testis
and
epididymis
weight,
sperm
morphology,
density
and
motility.
The
female
reproductive
parameters
evaluated
included
mating
index,
pregnancy
index,
fertility
index,
gestation
index,
number
of
live
births,
number
of
resorptions,
implants
per
litter,
corpora
lutea
and
pre­
and
post­
implantation
loss.
Concentrations
of
0,
100,
700
and
1300
ppm
bromodichloromethane
were
selected
for
use
in
this
study
based
on
decreased
water
consumption
observed
in
a
preliminary
14­
day
range­
finding
study
(
see
Section
V.
B.
1).
Two
groups
of
male
Sprague­
Dawley
rats
and
three
groups
of
female
Sprague­
Dawley
rats
were
assigned
to
treatment
groups
as
indicated
in
Table
V­
6.

Table
V­
6
NTP
(
1998)
Study
Design
Gender
Group
Description
#
Animals
per
Dose
Group
0
ppm*
100
ppm
700
ppm
1300
ppm
Male
A
non­
BrdU
treated
10
10
10
10
B
BrdU
treated
5
5
5
8
Female
A
peri­
conception
exposure
10
10
10
10
B
gestational
exposure
13
13
13
13
C
BrdU
treated,
periconception
exposure
5
5
5
8
*
Control
animals
received
deionized
water
Test
animals
were
dosed
for
25
to
30
days,
with
the
exception
of
Group
B
females
which
were
dosed
from
GD
6
to
evidence
of
littering/
birth
(
total
duration
approximately
15
to
16
days).
Based
on
measured
water
consumption,
the
nominal
concentrations
of
0,
100,
700
and
1300
ppm
were
equivalent
to
doses
of
0,
8,
41,
and
68
mg
bromodichloromethane/
kg­
day
for
all
male
rats
and
0,
14,
72
and
116
mg
bromodichloromethane/
kg­
day
for
all
female
rats
in
groups
A
and
C.
The
calculated
doses
for
Group
B
females
were
0,
13,
54,
and
90
mg/
kg­
day.
All
animals
survived
the
treatment
period,
with
the
exception
of
one
Group
A
male
in
the
700
ppm
dose
group.
Body
weight
and
food
and
water
consumption
were
decreased
at
many
time
points
for
animals
dosed
with
700
and
1300
ppm
bromodichloromethane.
Body
weights
in
the
dosed
groups
were
decreased
from
5%
to
13%,
food
consumption
was
decreased
from
14%
to
53%,
and
water
consumption
was
decreased
from
7%
to
86%
relative
to
control
animals.
However,
bromodichloromethane
exposure
did
not
alter
any
reproductive
parameter
investigated
in
males
or
females,
with
the
exception
of
a
non­
dose­
related
increase
in
the
number
of
live
fetuses
per
birth
Draft
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February
20,
2003
V
­
47
at
the
100
ppm
concentration
in
Group
C
females,
and
a
slight
decrease
in
the
number
of
corpora
lutea
at
the
700
ppm
concentration
in
Group
A
females.
On
the
basis
of
these
results,
NTP
(
1998)
concluded
that
bromodichloromethane
was
not
a
short­
term
developmental
or
reproductive
toxicant
any
of
the
doses
tested
in
the
study.
The
reproductive/
developmental
NOAELs
are
68
and
116
mg/
kg­
day
for
male
and
female
rats,
respectively.
The
adult
NOAEL
and
LOAEL
for
this
study
were
identified
on
the
basis
of
hepatic
effects,
which
are
discussed
in
detail
in
Section
V.
B.
1.

Bielmeier
et
al.
(
2001)
conducted
a
series
of
experiments
to
investigate
the
mode
of
action
in
bromodichloromethane­
induced
full
litter
resorption
(
FLR).
These
experiments
are
summarized
in
Table
V­
7.
This
series
of
experiments
included
a
strain
comparison
of
F344
and
Sprague­
Dawley
(
SD)
rats,
a
critical
period
study,
and
two
hormone
profile
studies.
In
the
strain
comparison
experiment,
female
SD
rats
(
13
to
14/
dose
group)
were
dosed
with
0,
75,
or
100
mg/
kg­
day
by
aqueous
gavage
in
10%
Emulphor
®
on
GD
6
to
10.
F344
rats
(
12
to
14/
dose
group)
were
concurrently
dosed
with
0
or
75
mg/
kg­
day
administered
in
the
same
vehicle.
The
incidence
of
FLR
in
the
bromodichloromethane­
treated
F344
rats
was
62%,
while
the
incidence
of
FLR
in
SD
rats
treated
with
75
or
100
mg/
kg­
day
of
bromodichloromethane
was
0%.
Both
strains
of
rats
showed
similar
signs
of
maternal
toxicity,
and
the
percent
body
weight
loss
after
the
first
day
of
dosing
was
comparable
for
SD
rats
(
no
resorption
observed)
and
the
F344
rats
that
resorbed
their
litters.
F344
rats
that
maintained
their
pregnancies
generally
did
not
lose
weight
after
the
first
dose,
although
they
did
experience
significantly
less
weight
gain
than
the
controls.
Both
strains
of
rats
had
similar
incidences
of
piloerection.
However,
the
strains
showed
different
ocular
responses
to
compound
administration.
One
half
(
7/
14)
of
the
treated
F344
rats
showed
lacrimation
and/
or
excessive
blinking
shortly
after
dosing
during
the
first
two
days
of
compound
administration.
In
comparison,
only
1/
28
of
the
SD
rats
exhibited
this
response.
The
study
authors
reported
that
lacrimation
was
not
predictive
of
FLR
in
F344
rats.
The
rats
were
allowed
to
deliver
and
pups
were
examined
on
postnatal
days
1
and
6.
Surviving
litters
appeared
normal
and
no
effect
on
post­
natal
survival,
litter
size,
or
pup
weight
was
observed.

Bielmeier
et
al.
(
2001)
conducted
a
second
experiment
to
identify
the
critical
period
for
bromodichloromethane­
induced
FLR.
Two
different
five
day
periods
during
organogenesis
were
compared.
F344
rats
(
12
to
13/
dose
group)
were
dosed
with
75
mg/
kg­
day
by
gavage
in
10%
Emulphor
®
on
GD
6
to
10
(
which
includes
the
luteinizing
hormone­
dependent
period
of
pregnancy)
or
GD
11
to
15
(
a
luteinizing
hormone­
independent
period).
Rats
(
8
to
10/
dose
group)
dosed
with
0
or
75
mg/
kg­
day
on
GD
6
to
15
served
as
negative
and
positive
controls,
respectively.
FLR
occurred
only
in
rats
treated
on
GD
6
to
10
or
GD
6
to
15
(
incidences
of
75%
and
50%,
respectively).
In
contrast,
all
rats
treated
with
bromodichloromethane
on
GD
11
to
15
maintained
their
litters.
Surviving
litters
appeared
normal
and
no
effect
on
post­
natal
survival,
litter
size,
or
pup
weight
was
observed.
This
finding
was
interpreted
by
the
study
authors
as
evidence
for
an
effect
of
bromodichloromethane
on
luteinizing
hormone
secretion
or
signal
transduction.

To
investigate
possible
endocrine
responses
to
bromodichloromethane
treatment
that
might
be
associated
with
FLR,
Bielmeier
et
al.
(
2001)
characterized
the
serum
profiles
of
Draft
­
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or
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February
20,
2003
V
­
48
luteinizing
hormone
(
LH)
and
progesterone
in
two
experiments.
Progesterone
is
necessary
for
the
maintenance
of
pregnancy,
while
LH
participates
in
the
maintenance
of
the
corpora
lutea
which
secrete
progesterone.
In
the
first
experiment,
F344
rats
(
8­
10/
dose)
were
given
a
100
Table
V­
7
Summary
of
Experiments
Conducted
by
Bielmeier
et
al.
(
2001)

Study/
Strain
Dose
(
mg/
kg­
day)
Treatment
Period
Number
of
animals
%
FLR
Treated
Pregnant
Resorbed
Strain
Comparison
F344
0
GD
6­
10
12
11
0
0
F344
75
GD
6­
10
14
13
8
62**

SD
0
GD
6­
10
13
13
0
0
SD
75
GD
6­
10
14
14
0
0
SD
100
GD
6­
10
14
14
0
0
Critical
Study
Period
F344
0
GD
6­
15
8
8
0
0
F344
75
GD
6­
15
10
10
5
50*

F344
75
GD
6­
10
12
12
9
75**

F344
75
GD
11­
15
13
13
0
0
Hormone
Profile
Ia
F344
0
GD
8­
9
8
7
0
0
F344
100
GD
8
10
10
6
60*

F344
100
GD
9
10
9
9
100***

Hormone
Profile
IIb
F344
0
GD
9
8
8
0
0
F344
75
GD
9
11
11
7
64*

F344
100
GD
9
10
10
9
90***

Source:
Table
1
in
Bielmeier
et
al.
(
2001)
Abbreviations:
GD,
gestation
day;
FLR,
full
litter
resorption;
SD,
Sprague­
Dawley
a
Tail
blood
collected
once
daily
on
GD
9
to
12.
b
Tail
blood
collected
at
0,
6,
12,
and
24
hours
after
dosing.
*
p<
0.05;
**
p<
0.01;
***
p<
0.001for
significant
differences
from
controls
(
Fisher's
Exact
Test).
Draft
­
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or
Quote
February
20,
2003
V
­
49
mg/
kg
dose
by
gavage
in
10%
Emulphor
®
on
GD
8
or
9.
Tail
blood
samples
were
collected
once
daily
on
GD
9
through
12
and
progesterone
and
LH
were
determined
by
ELISA
and
radioimmunoassay,
respectively.
As
reported
by
the
study
authors,
the
first
blood
collection
(
GD
9)
established
baseline
levels
for
the
GD
9­
treated
group
and
24­
hour
post­
dosing
levels
for
animals
treated
on
GD
8.
FLR
occurred
in
0,
60,
and
100%
of
the
control,
GD
8
and
GD
9
dose
groups,
respectively.
When
progesterone
was
measured
24
hours
after
dosing,
all
rats
that
resorbed
their
litters
had
markedly
reduced
levels
of
progesterone
when
compared
to
the
control
rats
or
to
rats
that
retained
their
litters.
The
mean
progesterone
(
±
standard
error)
levels
in
animals
dosed
on
GD
9
(
n
=
9)
that
resorbed
their
litters
decreased
from
a
baseline
level
of
137.94
±
11.44
ng/
mL
to
48.45
±
23.57
ng/
mL
within
24
hours.
The
mean
progesterone
levels
24
hours
post­
treatment
for
animals
treated
on
GD
8
was
67.01
±
16.22
ng/
mL
for
the
animals
that
resorbed
their
litters
(
n
=
6),
while
the
corresponding
control
group
mean
was
127.19
±
14.89
ng/
mL
(
n=
7).
Three
days
after
dosing
the
progesterone
levels
in
the
resorbed
groups
were
reported
to
be
comparable
to
those
of
two
nonpregnant
animals
in
the
study.
In
contrast
to
the
changes
in
progesterone
levels,
LH
levels
were
unaffected
in
treated
rats
when
measured
24
hours
after
dosing.
Serum
LH
levels
were
elevated
on
GD
11
to
12
in
the
resorbed
groups.
In
the
groups
of
animals
that
resorbed
their
litters,
the
mean
serum
levels
of
LH
rose
from
about
0.20
ng/
mL
on
GD
10
to
about
0.80
ng/
mL
on
GD
11
and
remained
elevated
on
GD
12.
These
levels
were
reported
to
be
comparable
to
the
levels
observed
in
two
nonpregnant
animals
in
the
study.
During
the
same
time
period,
the
control
levels
dropped
from
0.31
ng/
mL
to
0.14
ng/
mL.
Litters
delivered
to
dams
exposed
to
bromodichloromethane
had
significantly
heavier
pups
and
fewer
pups
than
control
litters.

In
the
second
hormone
profile
study
conducted
by
Bielmeier
et
al.
(
2001),
doses
of
0,
75,
or
100
mg/
kg­
day
were
administered
to
rats
(
8­
11/
dose)
on
GD
9
by
gavage
in
10%
Emulphor
®
.
Blood
samples
were
collected
at
0,
6,
12,
and
24
hours
after
dosing.
The
dose­
related
incidence
of
FLR
was
0%
(
0/
8),
64%
(
7/
11),
and
90%
(
9/
10),
respectively.
The
progesterone
levels
in
all
groups
peaked
at
six
hours.
The
peak
level
observed
in
dams
administered
75
mg/
kg
that
resorbed
their
litters
was
significantly
reduced
when
compared
to
the
controls.
The
peak
in
animals
receiving
100
mg/
kg
was
slightly
reduced,
but
the
effect
did
not
reach
statistical
significance.
At
12
hours,
the
progesterone
levels
of
dams
that
resorbed
their
litters
were
significantly
reduced
in
both
bromodichloromethane
dose
groups.
At
24
hours,
progesterone
levels
were
further
reduced
in
dams
that
displayed
FLR.
The
mean
progesterone
levels
in
dams
that
retained
their
litters
were
comparable
to
the
controls.
LH
levels
were
comparable
in
the
control
and
treatment
groups
at
all
tested
time
points.
When
analyzed
by
a
repeated
measures
procedure,
the
results
indicate
a
significant
decline
among
all
groups
over
the
24
hour
period.
However,
no
significant
differences
were
noted
between
groups.
In
contrast
to
the
results
for
pup
number
in
the
previous
hormone
profile
experiment,
litters
of
treated
dams
in
this
experiment
had
significantly
more
pups
than
controls.

The
series
of
experiments
conducted
by
Bielmeier
et
al.
(
2001)
identified
a
LOAEL
of
75
mg/
kg­
day
(
the
lowest
dose
tested)
based
on
FLR
in
F344
rats.
A
NOAEL
was
not
identified.
Draft
­
Do
Not
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or
Quote
February
20,
2003
V
­
50
The
Chlorine
Chemistry
Council
sponsored
a
range
finding
reproductive
toxicity
study
of
bromodichloromethane
in
rats
(
CCC,
2000c),
which
was
conducted
according
to
standard
U.
S.
EPA
test
guidelines
(
U.
S.
EPA,
1998c)
and
GLP
standards.
This
study
is
summarized
in
Christian
(
2001b).
Male
and
female
Sprague
Dawley
rats
(
10/
sex/
group)
were
randomly
assigned
to
five
exposure
groups.
Additional
rats
(
6
males/
group
and
15
females/
group)
were
assigned
to
satellite
groups
for
collection
of
samples
for
analysis
of
bromodichloromethane
concentrations
in
selected
tissues
and
fluids
(
see
Section
III.
B).
Bromodichloromethane
was
administered
to
parental
rats
(
P
generation)
in
drinking
water
at
concentrations
of
0,
50,
150,
450,
or
1350
ppm.
Exposure
began
14
days
before
cohabitation
and
continued
until
the
day
of
sacrifice.
Female
estrous
cycle
evaluations
were
performed
daily,
beginning
14
days
before
exposure
initiation
and
continuing
for
14
days
after
the
first
day
of
exposure.
Clinical
observations
were
recorded
daily
during
the
exposure
period.
Male
body
weights
were
recorded
weekly
during
the
entire
exposure
period
and
at
sacrifice;
female
body
weights
were
recorded
weekly
during
precohabitation
and
cohabitation,
on
GD
0,
7,
14,
21,
and
25,
and
on
lactation
days
(
LD)
1,
5,
8,
11,
15,
22,
and
29.
Lactation
was
extended
for
one
week
(
LD
22­
29)
beyond
the
normal
3­
week
period
because
F
1
pup
body
weights
in
the
three
highest
dose
groups
were
significantly
reduced
on
LD
21
relative
to
control
values
(
results
are
described
below).
Water
and
feed
consumption
were
recorded
weekly
and
at
sacrifice
for
males
during
the
entire
exposure
period
(
except
for
feed
consumption
during
cohabitation),
and
more
frequently
for
females
during
gestation
and
lactation.
On
LD
29,
two
F
1
pups
per
sex
were
selected
from
each
litter
for
an
additional
week
of
postweaning
observation,
provided
ad
libitum
access
to
water
containing
the
same
concentration
of
bromodichloromethane
administered
to
their
parents
(
P
generation),
and
sacrificed
on
Day
8
postweaning.
P
generation
female
rats
were
assessed
for
duration
of
gestation,
fertility
index,
gestation
index,
number
and
sex
of
offspring
per
litter,
number
of
implantation
sites,
and
clinical
signs
of
toxicity
during
the
postpartum
period.
During
lactation,
maternal
behavior
was
observed
and
recorded
on
LD
1,
5,
8,
11,
22,
and
29.
Litters
were
externally
examined
following
delivery
to
identify
the
number
and
sex
of
pups,
stillbirths
and
live
births,
and
gross
external
malformations.
Litters
were
observed
at
least
twice
daily
during
the
preweaning
and
postweaning
period
for
pup
deaths
and
clinical
signs
of
toxicity.
Litter
size
and
viability,
viability
indices,
lactation
indices,
percent
survival,
and
sex
ratios
were
calculated.
During
the
postweaning
period
of
observation,
body
weights
and
feed
consumption
were
recorded
at
weaning
and
on
day
8
postweaning;
water
consumption
was
recorded
daily.

At
the
end
of
the
parental
exposure
periods
(
64
days
for
males
and
a
maximum
of
74
days
for
females),
all
P
generation
rats
were
sacrificed
and
a
gross
necropsy
of
the
thoracic,
abdominal,
and
pelvic
viscera
was
performed.
In
addition,
testes
and
epididymides
were
excised
from
males
and
paired
organ
weights
were
measured.
F
1
pups
exposed
to
bromodichloromethane
in
their
drinking
water
for
one
week
following
weaning
were
sacrificed
on
Day
8
postweaning
and
examined
for
gross
lesions.
No
histopathology
was
performed
on
either
the
P
or
F
1
generation.

The
consumption
of
bromodichloromethane
was
calculated
from
measured
water
intake
and
measured
concentrations
of
the
test
article.
Mean
consumed
dosages
of
bromodichloromethane
for
P
generation
male
rats
during
the
entire
exposure
period,
P
generation
female
rats
during
different
physiologic
stages,
and
F
1
postweaning
rats
are
summarized
in
Table
V­
8.
Males
Draft
­
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or
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February
20,
2003
V
­
51
and
nonpregnant
female
rats
tended
to
consume
similar
amounts
of
bromodichloromethane.
Progressively
higher
dosages
were
consumed
by
female
rats
in
the
pre­
mating,
gestation,
and
lactation
periods,
respectively.
The
highest
dosages
among
all
groups
were
consumed
by
F
1
female
rats
during
the
one­
week
postweaning
observation
period.
A
possible
source
of
error
in
the
estimates
for
lactating
females
was
consumption
of
the
dams'
drinking
water
by
their
pups.

In
the
P
generation,
all
male
rats
and
all
females
except
one
survived
to
scheduled
sacrifice.
Exposure­
dependent
reductions
in
both
absolute
(
g/
day)
and
relative
(
g/
kg
body
weight­
day)
water
consumption
were
observed
in
all
rats
of
both
sexes
and
were
attributed
to
Table
V­
8
Mean
Consumed
Doses
(
mg/
kg­
day)
of
Bromodichloromethane
in
the
Range
Finding
Study
Conducted
by
CCC
(
2000c)

Gen.
Sex
Exposure
Interval
0
ppm
50
ppm
150
ppm
450
ppm
1350
ppm
P
M
Full
study
Study
Days
1­
64
0.0
4.2
±
0.4
11.8
±
1.8
27.5
±
3.4
67.2
±
5.6
P
F
Pre­
mating
Study
days
1­
15
0.0
4.7
±
0.8
13.3
±
2.0
23.5
±
5.3
70.8
±
1.8
P
F
Gestation
days
0­
21
0.0
5.4
±
0.7
16.3
±
2.2
41.7
±
6.4
111.7
±
6.2
P
F
Lactation
days
1­
15
0.0
11.0
±
1.9
31.4
±
2.6
90.3
±
7.3
222.4
±
19.9
F1
M
Postweaning
days
1­
8
0.0
13.6
±
3.5
41.4
±
7.1
106.9
±
20.8
297.8
±
113.8
F1
F
Postweaning
days
1­
8
0.0
13.9
±
2.6
40.1
±
6.8
117.9
±
42.7
333.6
±
110.6
taste
aversion.
Reduced
water
consumption
was
most
pronounced
during
the
first
week
of
exposure,
and
was
evident
during
precohabitation
and
cohabitation
in
both
sexes,
and
during
postcohabitation
in
males
and
gestation
in
females.
However,
the
decrease
in
water
consumption
during
these
times
was
not
as
severe
as
that
observed
during
the
first
week
of
exposure.
Decreased
water
consumption
was
not
clearly
noted
in
females
during
lactation,
presumably
reflecting
the
physiologic
demands
for
high
fluid
consumption
during
this
period.
Exposurerelated
decreases
in
feed
consumption
were
noted
for
males
and
females
in
the
150,
450,
and
1350
ppm
exposure
groups,
and
persisted
in
the
450
and
1350
ppm
females
during
gestation
and
lactation.
Treatment­
related
clinical
signs
of
toxicity
were
observed
in
both
sexes
in
the
1350
ppm
exposure
groups
and
were
considered
to
be
generally
associated
with
reduced
water
consumption.
Males
exhibited
dehydration,
emaciation,
chromorhinorrhea,
and
chromodacryorrhea
during
the
pre­
mating,
cohabitation
and
post­
cohabitation
periods;
however,
the
most
severe
symptoms
resolved
within
the
first
17
days
of
exposure.
Among
females,
urinestained
fur
was
observed
in
one
or
more
animals
in
the
three
highest
dose
groups
during
lactation
and
was
considered
to
be
treatment­
related.
Reductions
in
mean
body
weight
gain
and
body
weight
were
observed
in
male
rats
in
the
450
and
1350
ppm
exposure
groups
relative
to
controls.
Draft
­
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Not
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or
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February
20,
2003
V
­
52
These
effects
were
most
severe
during
the
first
week
of
exposure.
Mean
body
weight
gains
for
the
450
ppm
and
1350
ppm
male
groups
over
the
entire
exposure
period
were
91.3%
and
76.3%
of
the
control
values,
respectively.
At
study
termination,
mean
male
body
weights
were
96.5%
and
91.6
%
for
the
450
ppm
and
1350
ppm,
respectively,
relative
to
control
values.
In
female
rats,
reductions
in
body
weight
gain
and
body
weight
occurred
in
150,
450,
and
1350
ppm
groups.
These
effects
were
most
severe
during
the
first
week
of
exposure,
but
also
persisted
throughout
gestation
and
lactation.
During
gestation,
the
mean
reductions
in
female
body
weight
in
the
150,
450,
and
1350
ppm
groups
were
95.8%,
95.3%,
and
85.3%
of
the
control
values,
respectively.
Mean
body
weights
for
the
entire
lactation
period
were
not
presented
in
the
study
report;
however,
inspection
of
the
data,
presented
separately
for
LD
1,
8,
15,
22,
and
29,
indicated
that
female
body
weights
were
decreased
relative
to
controls
in
a
dose­
dependent
manner
in
the
three
highest
dose
groups
at
all
time
points.

No
gross
lesions
attributable
to
bromodichloromethane
were
observed
in
the
P
generation
male
or
female
rats
at
necropsy.
The
absolute
paired
epididymal
weights
were
slightly
reduced
(
93.2%
and
92.5%,
respectively)
in
the
450
and
1350
ppm
exposure
groups.
However,
relative
paired
epididymal
weights
were
unaffected,
suggesting
that
the
decreased
absolute
values
were
associated
with
the
reduced
terminal
body
weights
in
these
groups.
Absolute
and
relative
testes
weights
were
not
altered
by
exposure
to
bromodichloromethane.
No
effects
of
bromodichloromethane
were
observed
on
any
of
the
measured
reproductive
parameters
in
P
generation
male
or
female
rats.
However,
bromodichloromethane
exposure
was
associated
with
a
concentrationdependent
reduction
in
F
1
pup
body
weights
in
the
150,
450,
and
1350
ppm
exposure
groups.
Pup
weights
were
reported
for
postpartum
days
1,
5,
8,
15,
22,
and
29.
The
mean
litter
pup
weights
in
treated
groups
were
comparable
to
the
mean
litter
pup
weight
of
the
control
group
on
LD
1.
Beginning
on
LD
5,
reductions
in
mean
pup
weights
in
the
three
highest
dose
groups
increased
with
increasing
dose
and
duration
of
the
postpartum
period.
On
LD
29,
pup
weights
averaged
7,
12,
and
29%
less
than
controls
in
the
150,
450,
and
1350
ppm
exposure
groups,
respectively.
Reduced
body
weight
gain
continued
to
occur
in
the
F
1
pups
administered
parental
concentrations
of
bromodichloromethane
in
drinking
water
for
one
week
postweaning.
No
reductions
in
either
body
weight
gain
or
body
weight
were
observed
in
F
1
pup
litters
in
the
50
ppm
group
during
lactation
or
the
one­
week
postweaning
period.

Statistical
analysis
was
not
conducted
in
this
range
finding
study.
Based
on
decreased
pup
weight
and
pup
weight
gain,
the
LOAEL
for
developmental
toxicity
is
150
ppm,
and
the
corresponding
NOAEL
is
50
ppm.
Although
the
effect
of
reduced
water
consumption
on
the
decreases
in
feed
consumption,
body
weight
gain,
and
body
weight
observed
in
the
P
generation
adults
is
unclear,
the
LOAEL
for
parental
toxicity
is
considered
to
be
150
ppm
and
the
NOAEL
is
50
ppm.
Due
to
the
marked
changes
in
drinking
water
consumption
by
P
generation
female
rats
during
different
physiological
stages
(
pre­
mating,
mating,
gestation,
and
lactation),
it
is
not
possible
to
convert
the
administered
drinking
water
concentrations
into
biologically
meaningful
average
daily
doses.

The
Chlorine
Chemistry
Council
sponsored
a
developmental
toxicity
study
of
bromodichloromethane
in
rats
(
CCC,
2000d).
Data
from
this
study
are
summarized
in
Christian
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
53
et
al.
(
2001a).
This
study
was
conducted
in
accordance
with
U.
S.
EPA
Health
Effects
Test
Guidelines
OPPTS
870.3700:
Prenatal
Developmental
Toxicity
Study
(
U.
S.
EPA,
1998c)
and
U.
S.
EPA
Good
Laboratory
Practice
Standards
(
40
CFR
Part
160/
792).
Female
Sprague­
Dawley
rats
(
25/
exposure
group)
were
exposed
to
bromodichloromethane
in
the
drinking
water
at
concentrations
of
0,
50,
150,
450,
and
900
ppm
on
days
6
to
21
of
gestation
(
GD
6
to
21).
The
rats
were
examined
daily
during
the
exposure
period
for
clinical
signs
related
to
exposure,
abortions,
premature
deliveries
and
deaths.
Body
weights,
water
consumption,
and
feed
consumption
were
recorded
at
intervals
throughout
the
exposure
period.
All
study
animals
were
sacrificed
on
GD
21
and
caesarean­
sectioned.
A
gross
necropsy
of
the
thoracic,
abdominal,
and
pelvic
viscera
was
performed.
Data
was
collected
for
gravid
uterus
weight
(
with
cervix),
number
of
corpora
lutea/
per
ovary,
evidence
of
pregnancy,
number
and
distribution
of
implantation
sites,
live
and
dead
fetuses,
early
and
late
resorption,
and
placental
abnormalities
(
size,
color,
or
shape).
Individual
fetuses
were
weighed,
sexed,
and
examined
for
gross
external
abnormalities.
Approximately
one­
half
of
the
fetuses
in
each
litter
were
examined
for
soft
tissue
alterations
and
the
heads
of
these
fetuses
were
examined
by
free­
hand
sectioning.
The
remaining
fetuses
in
each
litter
were
examined
for
skeletal
alterations.

Consumed
dosages
for
GD
6
to
21
were
calculated
from
measured
water
consumption
and
measured
body
weights
and
averaged
0,
2.2,
18.4,
45.0,
and
82.0
mg/
kg­
day,
respectively.
No
abortions,
premature
deliveries,
deaths
or
treatment­
related
clinical
signs
were
observed
during
the
study
and
all
rats
survived
until
scheduled
sacrifice.
No
treatment­
related
gross
lesions
were
identified
at
autopsy.
Exposure­
related
decreases
in
maternal
body
weight
gains
occurred
in
all
groups
administered
bromodichloromethane
in
the
drinking
water
on
the
first
day
of
exposure
(
GD
6
to
7).
The
reduction
in
maternal
body
weight
gain
reached
statistical
significance
in
the
150,
450,
and
900
ppm
groups.
The
effect
was
most
severe
on
these
days
and
appeared
to
be
related
to
taste
aversion.
The
effect
on
maternal
body
weight
gain
was
persistent
in
the
450
and
900
ppm
exposure
groups.
In
contrast,
the
effect
was
transient
in
the
50
and
150
ppm
exposure
groups.
Average
body
weights
were
significantly
reduced
in
the
450
and
900
ppm
exposure
groups
on
GD
7
to
21.
Average
maternal
body
weights
in
the
same
groups
were
significantly
reduced
at
terminal
sacrifice
when
corrected
for
gravid
uterine
weight.

Statistically
significant,
exposure­
related
decreases
in
absolute
(
g/
day)
and
relative
(
g/
kgday
water
consumption
were
observed
in
all
groups
exposed
to
bromodichloromethane.
This
effect
was
evident
for
the
entire
exposure
period
(
GD
6
to
21)
and
the
entire
gestation
period
(
GD
0
to
21).
Within
the
exposure
period,
the
effects
were
most
pronounced
on
the
first
two
days
of
exposure
and
gradually
decreased
in
severity
with
continued
exposure.
Exposure­
related
decreases
in
absolute
and
relative
feed
consumption
were
observed
in
the
150,
450,
and
900
ppm
groups.
In
the
150
ppm
group,
the
effects
were
statistically
significant
only
on
GD
12
to
15
and
thus
were
considered
to
be
of
little
biological
importance
by
the
study
authors.
In
the
450
ppm
and
900
ppm
groups,
absolute
and
relative
feed
consumption
was
significantly
reduced
for
the
entire
exposure
period
(
GD
6
to
21),
the
entire
gestation
period
(
GD
0
to
21),
and
at
many
intervals
within
the
exposure
period.
The
effect
of
bromodichloromethane
on
feed
consumption
tended
to
be
most
severe
during
the
first
two
days
of
compound
administration.
Draft
­
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or
Quote
February
20,
2003
V
­
54
Caesarean
section
and
litter
parameters
were
unaffected
by
exposure
of
the
dams
to
bromodichloromethane
feed
concentrations
up
to
900
ppm.
Litter
averages
for
corpora
lutea,
implantations,
litter
sizes,
proportion
of
live
fetuses,
early
or
late
resorptions,
fetal
body
weights,
percent
reabsorbed
conceptuses,
and
percent
live
fetuses
were
comparable
among
all
study
groups
and
no
significant
differences
were
observed.
No
cases
of
full
litter
resorption
were
observed
and
there
were
no
dead
fetuses.
Late
resorption
occurred
in
one
control
group
litter.
All
placentae
appeared
normal.
All
values
for
the
examined
litter
parameters
were
within
the
historical
range
of
the
test
facility
(
Argus
Research
Laboratories,
Horsham,
PA)
or
litter
incidences
of
any
gross
external
or
soft
tissue
alterations.
With
respect
to
skeletal
alterations,
no
skeletal
malformations
were
observed
in
any
fetus.
The
only
statistically
significant
(
p<
0.01)
changes
in
the
occurrence
of
skeletal
variations
were
reversible
delays
in
ossification.
These
included
an
increased
fetal
incidence
(
fetal
incidence:
0
ppm,
1/
182;
50
ppm,
0/
199;
150
ppm,
0/
200;
450
ppm,
0/
188;
900
ppm,
4/
194;
litter
incidence:
0
ppm,
1/
23;
50
ppm,
0/
25;
150
ppm,
0/
25;
450
ppm,
0/
25;
900
ppm,
2/
25)
of
wavy
ribs
in
the
900
ppm
exposure
group
and
a
decreased
number
of
ossification
sites
per
fetus
per
litter
for
the
forelimb
phalanges
(
Mean
number
±
SD
of
ossification
sites:
8.14
±
0.91,
8.30
±
0.65,
8.09
±
0.63,
7.92
±
0.78,
7.46
±
0.78)
and
the
hindlimb
metatarsals
(
Mean
number
±
SD
of
ossification
sites:
4.81
±
0.25,
4.86
±
0.23,
4.78
±
0.27,
4.71
±
0.28,
4.53
±
0.33)
and
phalanges
(
Mean
number
±
SD
of
ossification
sites:
6.20
±
1.19,
6.20
±
1.17,
5.84
±
0.94,
5.86
±
0.79,
5.29
±
0.54).
The
increased
fetal
incidence
of
wavy
ribs
was
considered
unrelated
to
bromodichloromethane
exposure
by
the
study
authors
because
the
litter
incidence
(
the
more
relevant
measure
of
effect)
did
not
differ
significantly
from
the
control
and
was
within
the
historical
range
for
this
alteration
at
the
test
facility.

The
concentration­
based
maternal
NOAEL
and
LOAEL
for
this
study
were
150
ppm
and
450
ppm,
respectively,
based
on
statistically
significant,
persistent
reductions
in
maternal
body
weight
and
body
weight
gains.
Based
on
the
mean
consumed
dosage
of
bromodichloromethane,
these
concentrations
correspond
to
doses
of
18.4
mg/
kg­
day
and
45.0
mg/
kg­
day,
respectively.
The
concentration­
based
developmental
NOAEL
and
LOAEL
were
450
ppm
and
900
ppm,
respectively,
based
on
a
significantly
decreased
number
of
ossification
sites
per
fetus
for
the
forelimb
phalanges
and
the
hindlimb
metatarsals
and
phalanges.
These
concentrations
correspond
to
mean
consumed
doses
of
45.0
mg/
kg­
day
and
82.0
mg/
kg­
day,
respectively.

b.
Studies
in
Rabbits
The
Chlorine
Chemistry
council
sponsored
a
range­
finding
developmental
toxicity
study
in
New
Zealand
White
rabbits
(
CCC,
2000a).
The
data
from
this
study
have
been
summarized
in
Christian
et
al.
(
2001b).
This
study
was
conducted
in
accordance
with
U.
S.
EPA
Health
Effects
Test
Guidelines
OPPTS
870.3700:
Prenatal
Developmental
Toxicity
Study
(
U.
S.
EPA,
1998c)
and
U.
S.
EPA
Good
Laboratory
Practice
Standards
(
40
CFR
Part
160/
792).
Bromodichloromethane
was
provided
to
New
Zealand
White
presumed
pregnant
rabbits
(
5/
group)
in
the
drinking
water
at
concentrations
of
0,
50,
150,
450,
and
1350
ppm
on
GD
6
to
29.
Additional
rabbits
(
4/
group)
were
similarly
assigned
to
satellite
treatment
groups
for
use
in
the
collection
of
samples
for
analysis
of
tissue
concentrations
of
bromodichloromethane
(
discussed
in
Section
III.
B).
Body
weights
were
recorded
on
GDs
0,
4,
daily
during
the
exposure
Draft
­
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or
Quote
February
20,
2003
V
­
55
period,
and
on
the
day
of
sacrifice.
Feed
and
water
consumption
data
were
recorded
daily.
The
rabbits
were
sacrificed
on
GD
29
and
gross
necropsy
of
the
thoracic,
pelvic,
and
abdominal
viscera
were
performed.
The
gravid
uterus
was
excised
and
weighed.
Examinations
were
made
for
number
and
distribution
of
corpora
lutea,
implantation
sites,
early
and
late
resorptions
live
and
dead
fetuses.
Each
fetus
was
examined
for
gross
external
alterations
and
sex
(
by
internal
examination).

The
mean
consumed
daily
doses
of
bromodichloromethane
for
GDs
6
to
29
were
0.0,
4.9,
13.9,
32.3,
and
76.3
mg/
kg­
day,
as
determined
from
measured
body
weights
and
measured
water
consumption.
Absolute
(
g/
day)
and
relative
(
g/
kg­
day)
maternal
water
intake
for
the
exposure
period
was
decreased
in
each
group
administered
bromodichloromethane.
The
relative
consumption
values
were
92%,
87%,
67%,
and
53%
of
the
control
group
value,
respectively.
Absolute
and
relative
feed
consumption
values
were
reduced
in
a
time
(
onset
of
reductions
delayed
in
the
50
and
150
ppm
exposure
groups)
and
exposure­
dependent
manner.
The
relative
values
for
feed
consumption
were
96%,
96%,
90%,
and
82%
of
the
control
group
value
for
the
exposure
period.
No
deaths,
abortions,
or
premature
deliveries
occurred
during
the
study.
No
treatment­
related
clinical
signs
or
gross
lesions
were
observed.
Maternal
body
weight
gains
for
the
exposure
period
were
82%,
80%,
73%,
and
50%,
respectively,
relative
to
the
controls.
The
study
authors
questioned
whether
these
reductions
were
associated
with
bromodichloromethane
exposure
since
similar
changes
did
not
occur
in
the
satellite
exposure
group,
and
suggested
that
the
reduced
body
weight
gains
were
artifacts
of
the
small
sample
size
used
in
the
study.
When
body
weights
were
corrected
for
gravid
uterus
weight,
all
exposed
groups
in
the
main
study
experienced
body
weight
loss
while
body
weight
gain
occurred
in
the
control
group.
Absolute
uterine
weights
were
reduced
in
the
450
and
1350
ppm
groups.
This
finding
was
most
likely
associated
with
reduced
body
weight
in
these
groups,
since
relative
gravid
uterine
weights
in
all
dosed
groups
were
similar
to
that
of
the
control.

Litter
averages
for
corpora
lutea,
implantations,
litter
sizes,
live
and
dead
fetuses,
early
and
late
resorptions,
percent
dead
or
resorbed
conceptuses,
fetal
body
weights,
and
percent
live
male
fetuses
were
comparable
for
the
control
and
all
exposure
groups
and
within
the
historical
ranges
for
the
test
facility
(
Argus
Laboratories,
Horsham,
PA).
All
placentas
were
normal
in
appearance.
No
gross
external
fetal
alterations
were
observed
in
the
control
or
treatment
groups.
In
the
satellite
study
(
described
in
Section
III.
B),
analytical
analyses
detected
trace
amounts
of
bromodichloromethane
in
placental
samples
from
two
litters
in
the
1350
ppm
group
and
in
one
fetus
from
the
1350
ppm
group.
Bromodichloromethane
was
not
detected
in
amniotic
fluid
or
maternal
plasma.
One
litter
in
the
450
ppm
satellite
exposure
group
consisted
of
only
early
resorptions.
The
concentration­
based
LOAEL
for
maternal
toxicity
in
this
study
is
50
ppm,
the
lowest
concentration
tested,
based
on
reduced
body
weight
gain.
This
concentration
corresponds
to
a
mean
daily
dose
of
approximately
4.9
mg/
kg­
day.
The
concentration­
based
NOAEL
for
developmental
effects
was
1350
ppm
(
the
highest
dose
tested).
This
corresponds
to
a
mean
daily
intake
of
approximately
76.3
mg/
kg­
day.

The
Chlorine
Chemistry
Council
(
CCC,
2000b)
sponsored
a
developmental
toxicity
study
in
New
Zealand
White
rabbits.
Data
from
this
study
were
summarized
in
Christian
et
al.
(
2001a).
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
56
Bromodichloromethane
was
provided
to
pregnant
rabbits
(
25/
dose
group)
at
concentrations
of
0,
15,
150,
450,
and
900
ppm
in
the
drinking
water
on
GD
6­
29.
Consumed
doses
were
calculated
from
measured
water
intake
and
measured
body
weights
and
averaged
0,
1.4,
13.4,
35.6,
and
55.3
mg/
kg­
day,
respectively,
over
the
14
day
treatment
period.
Feed
consumption,
water
intake,
and
body
weight
were
monitored
daily
during
the
exposure
period.
The
rabbits
were
sacrificed
on
GD
29
and
examined
for
gross
lesions
of
the
thoracic,
abdominal,
and
pelvic
viscera.
Uterine
weight,
number
of
implantation
sites,
uterine
contents,
and
number
of
corpora
lutea
were
recorded.
Each
fetus
was
examined
for
weight,
gross
external
alterations,
skeletal
alterations,
and
sex.
Visceral
alterations
and
cavitated
organs
were
evaluated
by
dissection.
One
rabbit
in
the
900
ppm
dose
group
was
sacrificed
moribund
with
hindlimb
paralysis
caused
by
a
back
injury.
Another
rabbit
in
the
900
ppm
exposure
group
had
a
dead
litter
as
a
result
of
a
non­
treatment
related
uterine
abnormality.
No
treatment­
related
clinical
signs
or
necropsy
results
were
observed.
The
450
and
900
ppm
exposure
groups
had
significantly
reduced
feed
and
water
consumption
rates
throughout
the
exposure
period.
These
groups
also
had
significantly
reduced
body
weight
gains
and
corrected
(
for
weight
of
gravid
uterus)
body
weight
gains
for
both
the
bromodichloromethane
exposure
period
(
GD
6
to
29)
and
the
entire
gestation
period
(
GD
0
to
29).
Bromodichloromethane
had
no
observable
effect
on
implantations,
corpora
lutea,
live
litter
size,
early
or
late
resorptions,
percentage
of
male
fetuses,
percentage
of
resorbed
conceptuses,
or
fetal
body
weight.
The
number
of
litters
with
any
alteration,
the
number
of
fetuses
with
any
alteration,
the
average
percentage
of
fetuses
with
any
alteration
did
not
differ
significantly
from
the
control.
Although
statistically
significant
increases
in
the
number
of
fused
sterna
centra
were
observed
in
the
150
and
450
ppm
groups,
this
effect
was
not
dose­
related
and
the
observed
incidences
were
within
the
historical
range
for
the
testing
facility.
Litter
averages
for
ossification
sites
per
fetus
did
not
differ
significantly
from
the
control
and
were
within
historical
range
for
the
testing
facility.
The
NOAEL
and
LOAEL
identified
for
maternal
toxicity
in
this
study
were
13.4
mg/
kg­
day
(
150
ppm)
and
35.6
mg/
kg­
day
(
450
ppm),
respectively,
based
on
decreased
body
weight
gain.
The
developmental
NOAEL
was
55.3
mg/
kg­
day
(
900
ppm)
based
on
absence
of
statistically
significant,
dose­
related
effects
at
any
tested
concentration.

Christian
et
al.
(
2002)
summarized
the
results
of
a
two­
generation
reproductive
toxicity
study
on
bromodichloromethane
conducted
in
Sprague­
Dawley
rats.
The
study
was
sponsored
by
the
Chlorine
Chemistry
Council
(
CCC,
2002)
and
was
conducted
in
accordance
with
U.
S.
EPA
Health
Effects
Test
Guideline
OPPTS
870.3800:
Reproduction
and
Fertility
Effects
(
U.
S.
EPA,
1998b)
and
U.
S.
EPA
Good
Laboratory
Practice
Standards
(
40
CFR
Part
160/
792).
Bromodichloromethane
was
continuously
provided
to
test
animals
in
the
drinking
water
at
concentrations
of
0,
50,
150,
or
450
ppm.
Drinking
water
solutions
were
prepared
at
least
once
weekly
and
precautions
were
taken
to
prevent
contamination
of
the
solutions
by
extraneous
sources
of
chlorine.
Concentrations
were
verified
analytically
at
the
beginning
and
end
of
each
exposure
period.
The
tested
concentrations
were
selected
on
the
basis
of
results
obtained
in
the
developmental
toxicity
screening
study
conducted
by
NTP
(
1998)
and
data
obtained
in
a
range­
finding
study
(
CCC,
2001c;
Christian
et
al.,
2001b).
Exposure
of
the
parental
generation
(
30
rats/
sex/
concentration)
was
initiated
when
the
test
animals
were
approximately
43
days
of
age
and
continued
through
a
70­
day
pre­
mating
period
and
a
cohabitation
period
of
up
to
14
days.
Parental
generation
males
were
exposed
for
approximately
106
days
prior
to
sacrifice.
Exposure
Draft
­
Do
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or
Quote
February
20,
2003
V
­
57
of
parental
generation
female
rats
continued
through
gestation
and
lactation
for
a
total
exposure
period
of
approximately
118
days.
F1
generation
rats
were
exposed
to
bromodichloromethane
in
utero
and
by
consumption
of
the
dam's
drinking
water
during
the
lactation
period.
At
weaning,
F1
rats
(
30/
sex/
concentration)
were
selected
for
a
postweaning/
premating
exposure
period
of
at
least
64
days,
followed
by
a
cohabitation
period
of
up
to
14
days.
Exposure
continued
through
gestation
and
lactation.
F1
generation
females
delivered
litters
and
the
F2
litters
were
sacrificed
on
lactation
day
22.

During
the
course
of
the
experiment,
parental
and
F1
generation
rats
were
evaluated
for
viability,
clinical
signs,
water
and
feed
consumption,
and
body
weight.
Parental
and
F1
generation
females
were
evaluated
for
estrous
cycling
(
premating
and
during
cohabitation
until
mating
confirmed
and
at
sacrifice),
abortions,
premature
deliveries,
duration
of
gestation,
gestation
index,
fertility
index,
number
and
sex
of
offspring
per
litter,
general
postpartum
condition
of
dam
and
litter,
litter
size,
viability
index,
lactation
index,
percent
survival,
sex
ratio,
and
maternal
behavior.
Litters
were
examined
for
number
and
sex
of
pups,
stillbirths,
live
births,
and
gross
external
alterations.
F1
rats
selected
for
continued
evaluation
were
assessed
for
age
at
vaginal
patency
or
preputial
separation.
At
sacrifice,
test
animals
were
examined
for
gross
pathology,
organ
weights,
and
histopathology
(
control
and
high­
dose
groups,
10
parental
animals/
sex;
reproductive
organs
of
50
and
150
ppm
rats
suspected
of
reduced
fertility).
Male
rats
were
evaluated
for
sperm
concentration,
percent
motile
sperm,
sperm
morphology,
total
number
of
sperm,
and
testicular
spermatid
counts.
Females
were
evaluated
for
number
and
distribution
of
implantation
sites.
F1
weanlings
not
selected
for
continued
evaluation
(
3
pups/
sex/
litter,
when
available)
and
all
F2
weanling
rats
were
evaluated
for
gross
lesions,
terminal
body
weight,
and
organ
weights.

Key
findings
in
the
two­
generation
study
reported
by
CCC
(
2001c)
and
Christian
et
al.
(
2002)
include
the
following.
The
bromodichloromethane
dose­
equivalent
for
each
drinking
water
concentration
varied
by
sex
and
reproductive
status.
Average
daily
doses
estimated
for
the
50,
150
and
450
ppm
concentrations
were
4.1
to
12.6,
11.6
to
40.2,
and
29.5
to
109
mg/
kg­
day,
respectively,
as
calculated
by
the
study
authors.
One
death
in
the
150
ppm
group
and
three
deaths
(
including
one
humane
sacrifice)
in
the
450
ppm
group
were
associated
with
reduced
water
consumption,
weight
loss
and/
or
adverse
clinical
signs
and
may
have
been
compound­
related.
Adverse
clinical
signs
occurred
in
parental
generation
female
rats
and
F1
male
and
female
rats
in
the
150
and
450
ppm
exposure
groups.
Compound­
related
signs
included
chromorhinorrhea,
pale
extremities,
urine­
stained
abdominal
fur,
and
coldness
to
touch.
The
study
authors
attributed
these
signs
to
reduced
water
consumption.
Body
weight
and
body
weight
gain
were
significantly
reduced
in
the
450
ppm
parental
generation
males
and
females
and
150
and
450
ppm
F1
generation
males
and
females.
The
significantly
reduced
final
body
weight
in
450
ppm
parental
generation
females
was
associated
with
decreased
absolute
organ
weights
and
increased
relative
organ
weights
when
expressed
as
a
percentage
of
body
or
brain
weight.
Absolute
and
relative
water
consumption
rates
were
significantly
reduced
in
parental
and
F1
generation
males
and
females
at
all
concentrations
of
bromodichloromethane.
Water
intake
by
parental
and
F1
animals
was
generally
reduced
by
10
to
20
percent
in
the
150
and
450
ppm
groups
when
compared
to
the
controls.
Absolute
and
relative
feed
consumption
rates
were
reduced
in
males
and
females
of
both
generations
at
150
and
450
ppm
when
compared
with
the
controls.
There
were
no
gross
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February
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V
­
58
pathological
or
histopathological
indications
of
compound­
related
toxicity.
Most
indicators
of
reproductive
or
developmental
toxicity
examined
by
Christian
et
al.
(
2002)
were
not
significantly
affected
by
bromodichloromethane
treatment.
However,
F1
and
F2
generation
pup
body
weights
were
reduced
in
the
150
and
450
ppm
groups
during
the
lactation
period
after
the
pups
began
to
drink
the
water
provided
to
the
dams.
The
F1
generation
had
statistically
significant
reductions
in
pup
body
weight
at
weaning
on
lactation
day
22.
Reductions
in
F2
pup
body
weight
did
not
reach
statistical
significance.
Small
(

6%),
but
statistically
significant,
delays
in
F1
generation
sexual
maturation
occurred
at
150
(
males)
and
450
ppm
(
males
and
females)
as
determined
by
timing
of
vaginal
patency
or
preputial
separation.
The
study
authors
attributed
these
delays
to
significant
reductions
in
body
weight
at
weaning.
The
values
for
sexual
maturation
endpoints
in
the
150
and
450
ppm
exposure
groups
did
not
differ
significantly
from
control
values
when
body
weight
at
weaning
was
included
as
a
covariate
in
the
analysis.
Females
rats
with
vaginal
patency
not
evident
until
40
or
41
days
postpartum
(
i.
e.,
the
most
delayed)
in
the
150
and
450
ppm
groups
had
normal
estrus
cycles,
mated,
and
produced
litters.
Estrous
cycling
in
parental
generation
females
was
not
affected
by
exposure
to
bromodichloromethane.
A
marginal
effect
on
estrous
cyclicity
was
observed
in
F1
females
in
the
450
ppm
exposure
group.
This
effect
was
reported
to
be
associated
with
a
higher
incidence
of
rats
in
the
450
ppm
group
(
5/
30)
with
six
or
more
consecutive
days
of
diestrus
relative
to
the
controls
(
2/
30).
The
study
authors
considered
this
effect
to
be
a
secondary
response
associated
with
reduced
pup
weights
and
possible
inadvertent
stimulation
of
the
uterine
cervix
during
the
performance
of
vaginal
smears.
Averages
for
estrous
cycles
per
21
days,
cohabitation,
mating
indices,
and
fertility
indices
were
unaffected
by
exposure
to
bromodichloromethane.
Exposure
to
bromodichloromethane
had
no
effect
on
anogenital
distances
in
male
or
female
F2
pups.
The
results
of
this
study
appear
to
identify
NOAEL
and
LOAEL
values
for
reproductive
effects
of
50
ppm
(
4.1
to
12.6
mg/
kg­
day)
and
150
ppm
(
11.6
to
40.2
mg/
kg­
day),
respectively,
based
on
delayed
sexual
maturation.
However,
the
study
authors
have
questioned
whether
delayed
sexual
maturation
in
F1
males
associated
with
reduced
body
weight
should
be
treated
as
reproductive
toxicity
or
general
toxicity,
since
the
root
cause
appears
to
be
dehydration
brought
about
by
taste
aversion
to
the
compound.
The
parental
NOAEL
and
LOAEL
are
also
50
and
150
ppm,
respectively,
based
on
reduced
body
weight
and
body
weight
gain
in
F0
females
and
F1
males
and
females.

2.
Dibromochloromethane
Borzelleca
and
Carchman
(
1982)
evaluated
the
reproductive
toxicity
of
dibromochloromethane
in
a
two­
generation
study
with
ICR
Swiss
mice.
The
authors
used
a
modified
multigeneration
study
protocol
for
this
investigation.
Groups
of
10
males
and
30
females
(
F/
0
generation)
were
exposed
to
dibromochloromethane
in
drinking
water
at
concentrations
of
0,
0.1,
1.0,
or
4.0
mg/
mL
for
seven
weeks.
The
study
authors
did
not
estimate
average
daily
doses
for
all
treated
groups.
However,
they
did
indicate
that
the
highest
drinking
water
concentration
(
4.0
mg/
mL)
corresponded
to
an
average
daily
dose
of
685
mg/
kg­
day.
Using
this
conversion
factor,
drinking
water
concentrations
of
0.1
and
1.0
mg/
mL
dibromochloromethane
were
estimated
to
correspond
to
average
daily
doses
of
17
and
171
mg/
kg­
day,
respectively.
Following
the
initial
exposure
period,
the
F/
0
mice
were
mated
to
produce
the
F/
1a
litters.
Each
male
mouse
was
cohoused
for
seven
days
with
three
randomly
selected
females.
Two
weeks
after
weaning
of
the
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February
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2003
V
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59
F/
1a
litters,
the
F/
0
mice
were
randomly
re­
mated
to
produce
the
F/
1b
litters.
A
similar
protocol
was
followed
for
the
F/
1c
litters.
After
a
21­
day
postnatal
period,
the
F/
1a
and
F/
1c
litters
were
sacrificed
and
necropsied.
The
F/
1b
generation
was
culled.
The
surviving
males
and
females
(
10
males
and
30
females)
were
exposed
for
11
weeks
to
dibromochloromethane
in
drinking
water
at
concentrations
of
0,
0.1,
1.0
or
4.0
mg/
mL,
and
then
randomly
mated
to
produce
the
F/
2a
and
F/
2b
generations.
A
two­
week
interval
occurred
between
weaning
of
the
F/
2a
generation
and
remating
of
the
F/
1b
generation
to
produce
the
F/
2b
generation.
Thus,
parental
generations
(
F0
and
F/
1b)
were
exposed
continuously
to
dibromochloromethane
in
drinking
water
throughout
the
pre­
mating,
mating,
gestation,
and
lactation
periods
for
a
total
of
27
and
25
weeks,
respectively.
Following
weaning
of
their
final
litters,
both
parental
generations
were
sacrificed
and
necropsied.
The
F/
2a
and
F/
2b
generations
were
sacrificed
and
necropsied
following
a
21­
day
postpartum
survival
period.
Additionally,
a
selected
number
of
pups
from
the
final
matings
of
each
generation
(
i.
e.,
F/
1c
and
F/
2b)
were
screened
for
either
dominant
lethal
mutations
or
teratologic
abnormalities.

Body
weight
gain
and
drinking
water
consumption
were
recorded
weekly
and
semi­
weekly
for
the
F/
0
and
F/
1b
generations,
respectively.
Mating,
gestation,
gestation
survival,
and
lactation
survival
indices
were
calculated
for
each
mating.
During
the
21­
day
postpartum
period,
pups
were
counted
on
days
0,
4,
7,
14,
and
21,
and
sexed
on
days
7,
14,
and
21.
Viability
and
lactation
indices
were
calculated.
After
sacrifice
of
all
litters
except
F/
1b
on
day
21,
one
male
and
one
female
pup
per
litter
were
randomly
selected
for
necropsy.
For
teratology
screening,
treated
females
from
the
F/
0
and
F/
1b
generations
were
sacrificed
on
GD
18.
The
number
of
implantations,
resorptions,
and
live
and
dead
fetuses
were
counted.
Fetuses
were
individually
weighed
and
examined
for
gross
malformations;
a
selected
subset
of
fetuses
were
examined
for
either
skeletal
or
visceral
anomalies.
Statistical
analysis
was
conducted
on
all
endpoints,
using
parametric
or
nonparametric
procedures,
as
appropriate
for
different
endpoints.
For
statistical
analyses
performed
on
the
pups,
the
sampling
unit
was
the
litter.
Treatment­
related
effects
were
considered
to
be
statistically
significant
if
the
p
value

0.05.

As
compared
with
concurrent
controls,
final
body
weights
were
significantly
reduced
in
the
high­
dose
males
and
the
mid­
and
high­
dose
females
of
the
F/
0
and
F/
1b
generations.
Water
consumption
was
unaffected
by
treatment,
indicating
that
taste
aversion
was
not
a
factor
in
the
observed
decreases
in
body
weight.
Animals
in
both
the
F/
0
and
F/
1b
generations
exhibited
enlarged
livers
with
gross
morphological
changes,
interpreted
by
the
authors
as
indicative
of
hepatotoxicity.
The
incidence
and
the
severity
of
these
alterations
increased
with
increasing
dose,
with
0,
25,
70,
and
100%
of
the
F/
0
animals
and
0,
18,
64,
and
100%
of
the
F/
1b
animal
exhibiting
hepatic
discoloration,
fat
accumulation,
and/
or
lesions
at
0,
0.1,
1.0,
and
4.0
mg/
mL,
respectively.
Fertility
(
mating
index)
was
significantly
decreased
in
the
high­
dose
group
(
4.0
mg/
mL)
only
for
the
F/
2a
generation.
The
gestational
index
was
significantly
decreased
in
the
high­
dose
group
for
all
three
F/
1
generations.
Parental
ingestion
of
4.0
mg/
mL
dibromochloromethane
resulted
in
(
1)
decreased
litter
size
in
all
generations
(
F/
1a,
F/
1b,
F/
1c,
F/
2a,
and
F/
2b);
(
2)
decreased
viability
index
in
four
of
the
five
generations
(
F/
1a,
F/
1b,
F/
1c
and
F/
2a);
(
3)
decreased
lactation
index
in
the
F/
2b
generation;
and
(
4)
decreased
postnatal
body
weight
in
the
F/
2b
generation.
Parental
ingestion
of
1.0
mg/
mL
dibromochloromethane
produced
(
1)
decreased
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February
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2003
V
­
60
litter
size
in
the
F/
1c
generation;
(
2)
decreased
viability
index
in
the
F/
1b
generation;
(
3)
decreased
lactation
index
in
the
F/
1b
and
F/
2b
generations;
and
(
4)
decreased
postnatal
body
weight
in
the
F/
2b
generation.
The
only
statistically
significant
effect
observed
at
the
lowest
dose
tested
(
0.1
mg/
mL)
was
a
reduction
in
postnatal
body
weight
in
the
F/
2b
generation
on
PND
14;
this
effect
was
not
noted
on
PND
7
or
21.
No
statistically
significant
increases
in
dominant
lethal
or
teratogenic
effects
were
reported
in
either
the
F/
1c
or
F/
2b
generations.
Based
on
decreased
postnatal
body
weight
in
the
F/
2b
generation,
the
marginal
LOAEL
for
reproductive/
developmental
toxicity
is
17
mg/
kg­
day
and
a
NOAEL
could
not
be
determined.
The
developmental
LOAEL
is
considered
to
be
marginal
because
(
1)
this
effect
was
only
noted
in
one
of
the
two
litters
in
the
F/
2
generation;
(
2)
no
other
adverse
effects
were
observed
at
this
dose
level;
and
(
3)
it
was
unclear
from
the
report
how
many
litters
and
pups
per
litter
were
examined
for
postnatal
body
weight.
For
parental
toxicity,
liver
alterations
indicative
of
hepatotoxicity
were
clearly
evident
at
the
two
higher
doses
in
both
parental
generations.
At
the
lowest
dose
tested,
hepatic
changes
were
mainly
limited
to
discoloration,
presumably
due
to
the
accumulation
of
fat
deposits
(
Borzelleca
and
Carchman,
1982);
gross
morphology
was
normal,
and
histopathologic
examination
was
not
conducted.
Therefore,
the
adversity
of
this
effect
was
uncertain.
Based
on
these
considerations,
the
lowest
dose,
17
mg/
kg­
day,
is
considered
to
be
a
marginal
LOAEL
for
parental
toxicity
(
both
F0
and
F1/
b
generations)
and
a
NOAEL
could
not
be
determined.

Ruddick
et
al.
(
1983)
investigated
the
reproductive
and
developmental
toxicity
of
dibromochloromethane
in
Sprague­
Dawley
rats.
Pregnant
dams
(
10
to
12
animals
per
dose
group)
were
administered
gavage
doses
of
0,
50,
100,
or
200
mg/
kg­
day
in
corn
oil
on
GD
6­
15.
Body
weights
were
measured
on
GD
1,
on
GD
1
through
GD
15,
and
before
and
after
caesarean
section
on
GD
22.
On
GD
22,
females
were
anesthetized,
exsanguinated,
and
viscera
(
including
the
uteri)
were
examined
for
pathological
changes.
The
fetuses
were
removed,
weighed
individually,
and
examined
for
viability
and
external
malformations.
Two
pups
per
dam
were
placed
in
fixative
for
histopathological
examination.
Approximately
two­
thirds
of
the
remaining
live
fetuses
were
preserved
for
examination
for
skeletal
abnormalities.
The
remaining
fetuses
were
preserved
for
examination
for
visceral
alterations.
Maternal
blood
was
analyzed
for
standard
hematological
and
clinical
biochemistry
parameters.
Following
gross
pathological
examination
of
the
dams,
organ
weights
were
collected
for
liver,
heart,
brain,
spleen,
and
one
kidney.
Tissues
from
control
and
high
dose
dams
(
5
animals/
group)
were
subject
to
histopathological
examination.
Where
chemical
related
effects
were
observed,
the
affected
tissues
were
also
examined
in
the
mid­
dose
group.

Maternal
weight
gain
was
depressed
by
25%
in
the
high­
dose
group
relative
to
controls.
No
significant
effects
on
maternal
organ
weights,
hematology
and
clinical
chemistry,
number
of
resorption
sites,
number
of
fetuses
per
litter,
and
mean
fetal
body
weight
gain
were
observed
in
any
of
the
dose
groups.
No
treatment­
related
histopathology
was
noted
in
either
dams
or
fetuses.
There
were
no
skeletal
or
visceral
fetal
anomalies
attributed
to
dibromochloromethane
treatment.
Statistical
analysis
of
fetal
endpoints
was
not
conducted
by
the
study
authors.
However,
inspection
of
the
data
indicated
that
there
were
no
dose­
related
effects
(
e.
g.,
the
number
of
affected
fetuses/
number
of
affected
litters
for
sternebral
aberrations
was
3/
2,
2/
1,
1/
1,
and
1/
1
for
Draft
­
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or
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February
20,
2003
V
­
61
control,
low­,
mid­,
and
high­
dose
groups,
respectively).
The
power
of
this
experiment
was
limited
by
the
small
number
of
litters
per
dose
group.
In
the
absence
of
observed
fetal
effects,
the
NOAEL
for
developmental
toxicity
was
200
mg/
kg­
day,
the
highest
dose
tested,
and
a
LOAEL
could
not
be
determined.
Based
on
significantly
decreased
maternal
body
weight
gain,
the
LOAEL
and
NOAEL
for
maternal
toxicity
were
200
and
100
mg/
kg­
day,
respectively.

NTP
(
1996)
conducted
a
short­
term
reproductive
and
developmental
toxicity
screen
in
Sprague­
Dawley
rats.
Dibromochloromethane
was
administered
in
drinking
water
at
concentrations
of
0,
50,
150,
or
450
ppm
during
a
study
period
of
35
days.
Males
(
10/
group)
were
treated
from
study
days
6
through
34.
At
study
termination,
males
were
submitted
for
a
thorough
examination,
which
included
hematology,
clinical
chemistry,
gross
necropsy,
histopathology,
and
a
complete
sperm
evaluation
(
count,
density,
motility,
and
morphology).
Group
A
females
(
10/
group)
were
treated
from
study
days
1
through
34.
These
females
were
mated
to
treated
males
on
study
days
13
through
18
and
necropsied
on
study
day
34.
Group
B
females
(
13/
group)
were
mated
on
study
day
1
to
untreated
males,
treated
from
GD
6
through
parturition,
and
necropsied
on
postnatal
day
5.
No
hematology,
clinical
chemistry,
or
histopathology
was
conducted
on
the
females.

Based
on
measured
water
consumption,
the
authors
estimated
dose
levels
for
the
males
as
4.2,
12.4,
and
28.2
mg/
kg­
day,
for
Group
A
females
as
6.3,
17.4,
and
46.0
mg/
kg­
day,
and
for
Group
B
females
as
7.1,
20.0,
and
47.8
mg/
kg­
day.
A
few
changes
in
clinical
chemistry
were
noted
for
the
males.
Alkaline
phosphatase
and
5'
nucleotidase
were
increased
at
all
dose
levels
in
males,
but
reached
statistical
significance
only
at
the
low
dose
for
alkaline
phosphatase
and
at
the
mid
and
high
dose
for
5'
nucleotidase.
Total
serum
proteins
were
also
decreased
at
the
high
dose
in
males.
The
study
authors
noted
that
these
changes
could
reflect
mild
liver
damage.
However,
no
treatment­
related
microscopic
lesions
were
observed.
No
statistically
significant
effects
were
observed
on
any
sperm
parameter
investigated.
No
effect
was
observed
on
any
reproductive
or
fertility
measure
in
Group
A
or
B
females
at
any
dose.
The
proportion
of
male
pups
was
significantly
decreased
in
Group
B
females
at
the
high
dose
compared
to
the
control
value.
The
study
authors
did
not
consider
this
result
to
be
treatment­
related
because
the
control
value
(
0.61)
was
unusually
high
compared
to
historical
values,
and
the
result
for
the
high
dose
group
(
0.44)
was
within
historical
background.
Based
on
these
data,
the
authors
noted
that
dibromochloromethane
was
not
a
reproductive
toxicant
at
doses
up
to
the
high
dose
in
either
males
(
28.2
mg/
kgday
or
females
(
46.0
to
47.8
mg/
kg­
day).
Based
on
the
clinical
chemistry
changes,
the
authors
stated
that
administration
of
dibromochloromethane
may
have
resulted
in
general
toxicity
at
all
doses
in
the
male
treatment
groups.
The
observed
changes
in
clinical
chemistry,
however,
would
not
be
considered
adverse
for
the
following
reasons:
absence
of
clear
dose­
related
response,
small
magnitude
of
the
changes,
and
absence
of
supporting
histopathology
data.
Therefore,
this
study
identified
NOAEL
values
of
28.2
mg/
kg­
day
and
47.8
mg/
kg­
day
for
males
and
females,
respectively,
for
reproductive
and
systemic
effects.

3.
Bromoform
Draft
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February
20,
2003
V
­
62
Ruddick
et
al.
(
1983)
investigated
the
reproductive
and
developmental
toxicity
of
bromoform
in
Sprague­
Dawley
rats.
Pregnant
dams
(
14
to
15
animals/
dose
group)
were
administered
gavage
doses
of
0,
50,
100,
or
200
mg/
kg­
day
in
corn
oil
on
GD
6
to
15.
Body
weights
were
measured
on
GD
1,
on
GD
6
through
GD
15,
and
before
and
after
pups
were
delivered
by
caesarean
section
on
GD
22.
On
GD
22,
females
were
sacrificed
and
body
tissues
(
including
the
uterus)
were
removed
for
pathological
examination.
Females
were
evaluated
for
the
number
of
resorption
sites
and
the
number
of
fetuses.
Maternal
blood
samples
were
collected
and
evaluated
for
standard
hematology
and
clinical
chemistry
parameters.
The
liver,
heart,
brain,
spleen,
and
one
kidney
were
weighed.
Standard
histopathology
was
conducted
on
control
and
high
dose
females
(
5/
group).
All
fetuses
in
all
groups
were
individually
weighed,
and
evaluated
for
viability
and
external
malformations.
Histopathologic
examination
was
performed
on
two
pups
per
litter.
Of
the
remaining
live
fetuses,
approximately
two­
thirds
were
examined
for
skeletal
alterations
and
one­
third
for
visceral
abnormalities.

Maternal
weight
gain,
organ
weights,
hematology,
and
clinical
chemistry
were
unaffected
by
bromoform
treatment.
No
significant
differences
between
exposed
and
control
groups
were
observed
for
the
number
of
resorption
sites,
the
number
of
fetuses
per
litter,
fetal
weights,
fetal
gross
malformations,
and
visceral
abnormalities.
No
treatment­
related
histopathological
effects
were
noted
in
either
the
dams
or
fetuses.
However
an
elevation
in
the
incidence
of
skeletal
anomalies,
including
the
presence
of
a
14th
rib,
wavy
ribs,
and
interparietal
bone
deviations
was
reported
in
treated
animals.
An
increase
in
wavy
ribs
was
only
observed
in
the
high
dose
group.
The
number
of
affected
fetuses
/
number
of
affected
litters
for
the
presence
of
a
14th
rib
was
3/
3,
4/
3,
4/
3
and
7/
5
in
the
0,
50,
100,
and
200
mg/
kg­
day
groups,
respectively.
The
incidence
of
sternebral
aberrations
(
number
of
affected
fetuses/
number
of
affected
litters)
was
1/
1,
5/
3,
6/
5,
13/
8
in
the
0,
50,
100,
and
200
mg/
kg­
day
groups,
respectively.
Statistical
significance
for
fetotoxic
endpoints
was
not
reported
by
the
study
authors.
A
statistical
analysis
(
using
the
Fisher
Exact
test)
was
conducted
on
the
published
data
and
demonstrated
that
the
increase
in
sternebral
anomalies
was
significantly
different
from
controls
at
the
highest
dose
tested
(
200
mg/
kg­
day).
A
trend
test
showed
a
statistically
significant
dose­
related
trend
(
p<
0.002)
for
this
endpoint;
stepwise
analysis
indicated
that
the
trend
was
no
longer
significant
when
the
two
highest
doses
(
i.
e.,
200
and
100
mg/
kg­
day)
were
omitted
from
the
analysis.
These
findings
suggest
that
the
LOAEL
and
NOAEL
for
developmental
toxicity
were
100
and
50
mg/
kg­
day,
respectively.
In
the
absence
of
observed
maternal
effects,
the
NOAEL
for
maternal
toxicity
was
200
mg/
kg­
day,
and
a
LOAEL
could
not
be
determined.

NTP
(
1989b)
investigated
the
effect
of
bromoform
on
fertility
and
reproduction
in
Swiss
CD­
1
mice
using
a
continuous
breeding
protocol.
Twenty
male­
female
pairs
were
administered
daily
doses
of
50,
100,
or
200
mg/
kg­
day
by
gavage
in
corn
oil
and
forty
male­
female
pairs
were
dosed
with
the
corn
oil
vehicle
only.
Dose
selection
was
based
on
a
14­
day
range­
finding
study.
The
105­
day
dosing
period
included
a
seven­
day
precohabitation
phase
and
a
98­
day
cohabitation
phase.
The
parameters
evaluated
for
this
study
were
fertility,
litters
per
pair,
live
pups
per
litter,
proportion
of
pups
born
alive,
sex
of
live
pups,
or
pup
body
weights.
The
last
litter
born
(
generally
the
fifth
litter)
in
the
control
and
200
mg/
kg­
day
groups
during
a
holding
period
following
the
continuous
breeding
phase
were
reared
by
the
dams,
weaned
and
raised
to
sexual
Draft
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February
20,
2003
V
­
63
maturity
(
approximately
74
days)
while
receiving
the
same
treatment
(
vehicle
control
of
200
mg/
kg­
day
bromoform)
as
their
parents.
At
sexual
maturity,
males
and
females
from
different
litters
within
the
same
treatment
group
were
cohabited
for
seven
days
and
then
housed
individually
until
delivery.
The
endpoints
for
this
mating
trial
were
the
same
as
for
the
parental
generation.
At
study
termination,
the
F
1
mice
were
weighed,
necropsied
and
evaluated
for
selected
organ
weights,
epididymal
sperm
motility,
and
sperm
morphology.
Selected
organs
were
fixed
for
histopathological
examination.

In
the
200
mg/
kg­
day
treatment
group,
the
body
weights
of
dams
at
delivery
were
consistently
less
than
the
control
group
value.
The
reduction
in
body
weight
was
statistically
significant
after
delivery
of
the
first,
second,
fourth,
and
fifth
litters.
The
fertility
index
for
the
parental
generation
was
100%
for
the
control
and
treated
groups
(
a
breeding
pair
was
designated
as
fertile
if
they
produced
at
least
one
live
or
dead
pup).
There
was
no
detectable
effect
of
treatment
on
the
number
of
litters
per
pair,
the
number
of
live
pups
per
litter,
the
proportion
of
pups
born
alive,
the
sex
of
live
pups,
or
pup
body
weights.
The
gestational
period
was
similar
across
groups.
However,
postnatal
survival
of
F
1
pups
in
the
200
mg/
kg­
day
group
was
significantly
lower
than
in
the
control
group.
The
study
authors
reported
that
this
difference
was
primarily
attributable
to
three
dams
who
lost
all
of
their
pups
by
postnatal
day
4.
One
dam
in
the
control
group
also
lost
her
litter
by
postnatal
day
4.
The
study
authors
noted
that
this
result
is
consistent
with
a
treatment
effect
on
early
maternal
behavior,
early
lactational
failure,
and/
or
the
postnatal
developmental
processes.
When
F
1
mice
were
cohabited
for
one
week,
no
effect
of
treatment
on
mating
index
or
fertility
was
observed.
There
were
no
significant
differences
relative
to
control
values
for
the
number
of
live
pups
per
litter
(
male,
female,
or
combined),
the
proportion
of
live
pups,
the
proportion
of
male
pups,
or
pup
weight
at
birth.
At
sacrifice,
male
and
female
F
1
mice
administered
200
mg/
kg­
day
exhibited
increased
relative
liver
weights
and
decreased
relative
kidney
weights
as
compared
with
control
values.
The
mean
body
weight
for
F
1
males
was
significantly
less
than
the
mean
weight
of
the
male
control
group.
Histopathological
evaluation
revealed
minimal
to
moderate
hepatocellular
degeneration
in
the
livers
of
high­
dose
F
1
male
and
female
mice.
Bromoform
treatment
had
no
effect
on
epididymal
sperm
density,
motility,
or
morphology
in
F
1
males.
No
treatment­
related
histologic
lesions
were
observed
in
the
seminal
vesicles,
coagulating
glands,
or
prostate
glands
of
males,
or
in
the
lung,
kidney,
or
thyroid
gland
of
males
or
females.
Based
on
liver
histopathology,
decreased
postnatal
survival,
and
other
signs
of
toxicity
(
e.
g.,
increased
relative
liver
and
decreased
relative
kidney
weights)
in
F
1
mice
of
both
sexes
at
the
highest
dose
tested,
the
LOAEL
for
developmental
toxicity
is
200
mg/
kg­
day,
and
the
NOAEL
is
100
mg/
kg­
day.
Based
on
consistently
decreased
body
weights
of
pregnant
dams
at
delivery,
the
LOAEL
for
maternal
toxicity
is
200
mg/
kg­
day
and
the
NOAEL
is
100
mg/
kg­
day.

Table
V­
9
Summary
of
Reproductive
Studies
of
Brominated
Trihalomethanes
Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Bromodichloromethane
Table
V­
9
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
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or
Quote
February
20,
2003
V
­
64
Ruddick
et
al.
(
1983)
Rat
Sprague­
Dawley
Gavage
(
Corn
oil)
F
9­
14
GD
6­
15
0
50
100
(
NOAEL)
1
200
(
LOAEL)
Developmental
toxicity
study.
Statistically
decreased
maternal
wt.
gain
at
high
dose
(
38%);
increased
litter
incidence
of
sternebral
aberrations.
Statistical
significance
not
evaluated
for
fetotoxic
endpoints
by
study
authors;
statistical
analysis
conducted
on
published
data
for
fetal
effects.
Trend
test
indicated
statistical
effect
for
sternebral
anomalies
at
highest
dose
tested.
Maternal
LOAEL
and
NOAEL
are
200
and
100
mg/
kgday
respectively.

Klinefelter
et
al.
(
1995)
Rat
F344
Drinking
Water
M
7
52
weeks
0
22
(
NOAEL)
39
(
LOAEL)
Reproductive
toxicity
study.
Decreased
sperm
velocity
at
39
mg/
kgday
No
histopathological
alterations
noted
in
any
reproductive
tissue
examined.

Narotsky
et
al.
(
1997)
Rat
F344
Gavage
(
Corn
oil)
F
12­
14
GD
6­
15
0
25
(
NOAEL)
50
(
LOAEL)
75
Developmental
toxicity
study
comparing
the
use
of
different
gavage
dosing
vehicles.
Reduced
maternal
weight
gain
GD
6­
8
and
lacrimation
at
50
and
75
mg/
kg­
day.
Full
litter
resorption
(
FLR)
observed
at
50
and
75
mg/
kg­
day
(
8%
and
83%,
respectively).
No
effects
on
postnatal
survival,
pup
weight,
or
pup
survival
in
surviving
litters.
ED05
and
BMD
for
FLR
calculated
by
study
authors
as
33.3
and
11.3
mg/
kg­
day,
respectively.
Maternal
LOAEL
and
NOAEL
are
50
and
25
mg/
kg­
day,
respectively.

Narotsky
et
al.
(
1997)
Rat
F344
Gavage
(
Aqueous)
F
12­
14
GD
6­
15
0
25
(
NOAEL)
50
(
LOAEL)
75
Developmental
toxicity
study
comparing
the
use
of
different
gavage
dosing
vehicles.
Reduced
maternal
weight
gain
GD
6­
8
at
all
dose
levels.
Full
litter
resorption
(
FLR)
observed
at
50
and
75
mg/
kg­
day
(
17
and
21%,
respectively).
No
effects
observed
on
postnatal
survival,
pup
weight,
or
pup
survival
in
surviving
litters.
Maternal
LOAEL
is
25
mg/
kg­
day;
maternal
NOAEL
not
determined.

NTP
(
1998)
Rat
Sprague­
Dawley
Drinking
Water
M
5­
10
25­
30
days
0
8
41
68
(
NOAEL)
Reproductive/
developmental
toxicity
study.
Decreased
food
and
water
consumption;
decreased
body
weight.
No
dose­
related
changes
in
reproductive/
developmental
parameters
NTP
(
1998)
Rat
Sprague­
Dawley
Drinking
Water
F
5­
10
25­
30
days
0
14
72
116
(
NOAEL)
Reproductive/
developmental
toxicity
study.
Decreased
food
and
water
consumption;
decreased
body
weight.
No
dose­
related
changes
in
reproductive/
developmental
parameters
Table
V­
9
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
65
Bielmeier
et
al.
(
2001)
Rat
F344
Gavage
(
Aqueous)
F
8­
13
GD
6­
15
GD
6­
10
GD
11­
15
0
75
(
LOAEL)
Critical
exposure
period
study.
Full
litter
resorption
observed
in
animals
treated
on
GD
6­
10,
but
not
in
animals
treated
on
GD
11­
15.

Bielmeier
et
al.
(
2001)
Rat
F344/
Sprague­
Dawley
Gavage
(
Aqueous)
F
12­
14
GD
6­
10
F344
0
75
(
LOAEL)

Sprague­
Dawley
0
75
100
(
NOAEL)
Strain
comparison
study.
Full
litter
resorption
observed
in
F344
rats
but
not
in
concurrently
dosed
Sprague­
Dawley
rats.

Bielmeier
et
al.
(
2001)
Rat
F344
Gavage
(
Aqueous)
F
8­
11
GD
9
0
75
(
LOAEL)
100
Hormone
profile
study.
Full
litter
resorption
observed
at
both
doses.
Decreased
serum
progesterone
levels
in
F344
rats
which
experienced
FLR.

CCC
(
2000c)
Christian
et
al.
(
2001b)
Rat
Sprague­
Dawley
Drinking
Water
F,
M
10
64­
74
days
Males2
(
days
1­
64
i.
e.,
for
14
days
premating,
during
mating,
and
for

6
weeks
following
mating)

Females
2
(
days
1­
74,
i.
e.,
for
14
days
premating,
during
mating,
GD
0­
21,
lactation
days
1­
29)
0
ppm
50
ppm
(
NOAEL)
150
ppm
(
LOAEL)
450
ppm
1350
ppm
Range
finding
reproductive/
developmental
toxicity
study.
Decreased
body
weight
gain
and
terminal
body
weight
(>
10%)
in
males
at
highest
dost
tested
but
no
apparent
effects
on
reproductive
endpoints
at
any
dose.

Maternal
toxicity
(
reduced
body
weight
and
body
weight
gain
and
decreased
food
and
water
consumption)
at
150
ppm
and
higher.
Dose­
dependent
decreases
in
mean
pup
weight
gain
and
pup
weights
beginning
on
lactation
day
5­
29
in
3
highest
dose
groups.
Decreased
pup
body
weight
gain
and
body
weight
also
observed
in
3
highest
dose
groups
in
pups
treated
for
one
week
postweaning
at
parental
drinking
water
concentrations.

Reproductive/
developmental
LOAEL
and
NOAEL
are
150
and
50
ppm,
respectively,
based
on
decreased
pup
wt
and
wt
gain;
parental
LOAEL
and
NOAEL
are
150
and
50
ppm,
respectively,
based
on
reduced
body
wt
gain
and
wt.
Findings
confounded
by
effects
of
decreased
water
consumption
at
various
time
points
during
treatment.

CCC
(
2000d)
Christian
et
al.
(
2001a)
Rat
Sprague­
Dawley
Drinking
Water
F
25
GD
6­
21
0
2.2
18.4
45.0
(
NOAEL)
82.0
(
LOAEL)
Developmental
toxicity
study.
Decreased
maternal
body
weight
and
body
weight
gain
at
45.0
mg/
kg­
day.
Developmental
LOAEL
based
on
slightly
decreased
number
of
ossification
sites
in
the
hindlimb
(
metatarsals
and
phalanges)
and
forelimb
(
phalanges).
Maternal
LOAEL
and
NOAEL
are
82.0
and
45.0
mg/
kg­
day,
respectively.
Table
V­
9
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
66
CCC
(
2000a)
Christian
et
al.
(
2001b)
Rabbit
New
Zealand
White
Drinking
Water
F
5
GD
6­
29
0.0
4.9
13.9
32.3
76.3
(
NOAEL)
Range
finding
developmental
study.
Decreased
maternal
body
weight
gain
and
water
and
feed
consumption
at
all
tested
doses.
No
treatment­
related
changes
in
reproductive
or
developmental
endpoints.
Study
authors
considered
maternal
LOAEL
to
be
<
4.9
mg/
kg­
day,
based
on
significantly
reduced
body
weight
gain.

CCC
(
2000b)
Christian
et
al.
(
2001a)
Rabbit
New
Zealand
White
Drinking
Water
F
25
GD
6­
29
0
1.4
13.4
35.6
55.3
(
NOAEL)
Developmental
toxicity
study.
Reduced
maternal
weight
gain
at
35.3
mg/
kg­
day.
No
dose­
related
changes
in
reproductive
or
developmental
parameters.
Maternal
LOAEL
and
NOAEL
are
35.3
and
13.4
mg/
kg­
day,
respectively.

CCC
(
2002)
Rat
Sprague­
Dawley
Drinking
Water
M,
F
30
Up
to
118
days
Males
F
0:
106
d
(
incl.
70
d
pre­
mating)

F
1:
64
d
postweaning
14
d
cohab.

Females
F
0:
118
d
(
incl.
gest.,
lactation
F
1
64
d
postweaning
14
d
cohab.,
gest.,
lactation
0
50
(
NOAEL)
150
(
LOAEL)
450
Reproductive/
developmental
LOAEL
and
NOAEL
are
150
and
50
ppm,
respectively,
based
on
delayed
sexual
maturation
in
F
1
males;
parental
LOAEL
and
NOAEL
are
150
and
50
ppm,
respectively,
based
on
reduced
body
wt
gain
and
wt
in
F
0
females
and
F
1
males
and
females.
Findings
confounded
by
effects
of
decreased
water
consumption
as
a
result
of
taste
aversion
to
the
test
compound.

Dibromochloromethane
Borzelleca
and
Carchman
(
1982)
Mouse
ICR
Swiss
Drinking
Water
M
F
10
30
25­
27
weeks
0
17
(
marginal
LOAEL)
171
685
Multi­
generation
reproductive
toxicity
study.
Significant
high­
dose
effects
include
decreased
gestational
index
in
F1
generation
at
high
dose
and
decreased
litter
size
in
F1
and
F2
generations.
Significant
mid­
dose
effects
include
decreased
litter
size,
decreased
viability
index,
decreased
lactation
index,
and
decreased
postnatal
body
weight
in
some
F1
and/
or
F2
generations.
Only
significant
low­
dose
effect
is
reduced
postnatal
body
wt
in
F/
2b
generation
on
postnatal
day
14.
Hepatic
effects
observed
in
both
parental
generations
at
all
doses;
liver
effects
marginal
at
low
dose.
Parental
marginal
LOAEL
is
17
mg/
kg­
day;
parental
NOAEL
not
determined.
Table
V­
9
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
67
Ruddick
et
al.
(
1983)
Rat
Sprague­
Dawley
Gavage
(
Corn
oil)
F
10­
12
GD
6­
15
0
50
100
200
(
NOAEL)
Developmental
toxicity
study.
Significantly
depressed
maternal
wt.
gain
at
high
dose
(
25%);
increased
maternal
relative
liver
wt.
and
kidney
Statistical
significance
of
fetal
endpoints
not
evaluated
by
study
authors.
Based
on
data
inspection,
no
dose­
related
skeletal
or
visceral
effects
observed
in
litters.
Maternal
LOAEL
and
NOAEL
are
200
and
100
mg/
kgday
respectively.

NTP
(
1996)
Rat
Sprague­
Dawley
Drinking
Water
M
10
29
days
4.2
12.4
28.2
(
NOAEL)
Reproductive/
developmental
toxicity
study.
No
treatment­
related
effects
on
measured
sperm
parameters
NTP
(
1996)
Rat
Sprague­
Dawley
Drinking
Water
F
10
35
days
6.3
17.4
46.0
(
NOAEL)
Reproductive/
developmental
toxicity
study.
Exposure
occurred
during
a
6­
day
mating
period
and
most/
all
of
gestation.
No
clearly
adverse
effect
on
any
reproductive
or
developmental
endpoint
at
tested
doses
NTP
(
1996)
Rat
Sprague­
Dawley
Drinking
Water
F
13

16
days
(
GD
6
to
parturition)
7.1
20.0
47.8
(
NOAEL)
Reproductive/
developmental
toxicity
study.
No
clearly
adverse
effect
on
any
reproductive
or
developmental
endpoint
at
tested
doses
Table
V­
9
(
cont.)

Reference
Species
Route
Sex
Number
per
dose
group
Duration
Dose
(
mg/
kg­
day)
Results
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
68
Bromoform
Ruddick
et
al.
(
1983)
Rat
Sprague­
Dawley
Gavage
(
Corn
oil)
F
14­
15
GD
6­
15
0
50
(
NOAEL)
100
(
LOAEL)
200
Developmental
toxicity
study.
No
statistically
significant
maternal
effects.
Apparent
treatment­
related
increases
in
sternebral
aberrations
and
other
skeletal
endpoints.
Statistical
significance
of
fetotoxic
endpoints
not
evaluated
by
study
authors.
Statistical
analysis
conducted
on
published
data
indicated
significant
increase
in
sternebral
aberrations
at
two
highest
doses.
Maternal
NOAEL
is
200
mg/
kg­
day;
LOAEL
not
determined.

NTP
(
1989b)
Mouse
ICR
Swiss
Gavage
(
oil)
M
F
20
20
105
days
0
50
100
(
NOAEL)
200
(
LOAEL)
Continuous
breeding
reproductive
toxicity
protocol.
Maternal
body
weights
significantly
decreased
at
highest
dose
tested.
Decreased
postnatal
survival,
organ
wt.
changes,
and
liver
histopathology
observed
in
F1
mice
of
both
sexes
in
high­
dose
group.
No
noted
effects
on
fertility,
litters/
pair,
live
pups/
litter;
proportion
of
live
births,
sex
of
live
pups,
or
pup
body
weight.
Maternal
LOAEL
and
NOAEL
are
200
and
100
mg/
kg­
day,
respectively.

1
NOAEL
and
LOAEL
values
reported
in
this
column
are
for
developmental/
reproductive
toxicity
effects.
The
NOAEL
and
LOAEL
values
for
parental
toxicity
are
reported
in
the
"
Results"
column.
2
Doses
for
this
study
are
presented
as
ppm
in
drinking
water;
due
to
marked
changes
in
adult
female
water
consumption
during
different
physiologic
stages
(
i.
e.,
pre­
mating,
mating,
gestation,
and
lactation),
it
is
not
possible
to
convert
administered
drinking
water
concentrations
into
biologically
meaningful
average
daily
doses.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
69
F.
Mutagenicity
and
Genotoxicity
1.
Bromodichloromethane
The
results
of
in
vivo
and
in
vitro
tests
conducted
to
evaluate
the
mutagenicity,
genotoxicity
and
neoplastic
transformation
potential
of
bromodichloromethane
are
summarized
in
Table
V­
10
at
the
end
of
this
section.

In
Vitro
Assays
Simmon
and
Tardiff
(
1978)
reported
that
nonactivated
bromodichloromethane
was
mutagenic
in
S.
typhimurium
strain
TA100
when
assayed
in
a
desiccator
containing
the
test
compound
in
the
atmosphere.
The
minimum
amount
of
bromodichloromethane
required
to
elicit
a
mutagenic
response
following
addition
to
the
desiccator
was
600
µ
mol.

Ishidate
et
al.
(
1982)
assayed
the
mutagenicity
of
bromodichloromethane
in
S.
typhimurium
strain
TA100
in
the
presence
and
absence
of
rat
liver
S9
fraction.
Increased
mutation
frequencies
were
observed
only
in
the
absence
of
S9
activation.
In
contrast,
chromosomal
aberrations
in
Chinese
hamster
fibroblasts
were
observed
in
the
presence,
but
not
the
absence,
of
S9
fraction.
The
concentrations
tested
in
these
assays
were
not
reported.

Nestmann
and
Lee
(
1985)
investigated
the
mutagenicity
of
bromodichloromethane
at
12
to
1,200
µ
M
in
S.
cerevisiae
strains
D7
and
XV185­
14C
in
the
presence
or
absence
of
S9
activation.
No
clear
increase
in
convertants
or
in
revertants
of
strain
XV185­
14C
was
observed
for
bromodichloromethane
in
the
presence
or
absence
of
S9
activation.

NTP
(
1987)
reported
that
bromodichloromethane
was
not
mutagenic
when
tested
using
a
preincubation
protocol
in
S.
typhimurium
strains
TA1535,
TA1537,
TA98,
or
TA100
at
concentrations
reaching
cytotoxic
levels
(
20
µ
mol/
plate;
3,333
µ
g/
plate).
Testing
was
done
in
the
absence
of
S9
and
in
the
presence
of
S9
prepared
from
Aroclor­
induced
male
hamster
or
rat
liver.
NTP
concluded
that
the
negative
results
may
have
been
due
to
volatilization
of
the
test
compound
from
the
test
system.
Bromodichloromethane
was
not
mutagenic
in
the
mouse
lymphoma
L5178Y/
TK+/­
assay
in
the
absence
of
S9,
but
did
induce
dose­
related
increases
in
forward
mutations
at
S9­
activated
concentrations
greater
than
or
equal
to
2,000
µ
M
(
300
µ
g/
mL).
Cytogenetic
tests
with
Chinese
hamster
ovary
cells
(
CHO)
cells
were
reported
in
this
study
and
also
by
Anderson
et
al.
(
1990).
There
was
no
evidence
of
induction
of
chromosomal
aberrations
following
treatment
with
up
to
30,000
µ
M
(
5,000
µ
g/
mL)
in
either
the
presence
or
absence
of
exogenous
metabolic
activation.
There
was
also
no
evidence
of
sister
chromatid
exchanges
induced
by
the
nonactivated
material.
In
the
presence
of
S9
activation,
one
of
three
assays
resulted
in
a
positive
response
at
doses
greater
than
or
equal
to
24,400
µ
M
(
4,000
µ
g/
mL).
These
results
are
difficult
to
evaluate
because
cytotoxicity
was
observed
at
similar
concentrations
in
the
other
trials.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
70
Varma
et
al.
(
1988)
tested
bromodichloromethane
for
mutagenicity
in
S.
typhimurium
strains
TA1535,
TA1537,
TA98,
and
TA100.
In
the
absence
of
S9
fraction,
bromodichloromethane
at
nonactivated
concentrations
of
2.4
to
3.2
µ
mol/
plate
induced
mutations
in
strain
TA1537.
There
was
no
effect
in
the
other
strains.

Bromodichloromethane
was
positive
for
the
induction
of
DNA
damage
in
the
presence
and
absence
of
exogenous
activation,
based
on
the
results
of
the
SOS
chromotest
(
LeCurieux
et
al.,
1995).
Bromodichloromethane
gave
a
negative
result
in
the
fluctuation
test
modification
of
the
S.
typhimurium
reverse
mutation
assay.

Several
studies
have
evaluated
induction
of
sister
chromatid
exchanges
following
exposure
to
bromodichloromethane.
Morimoto
and
Koizumi
(
1983)
investigated
the
ability
of
bromodichloromethane
to
induce
sister
chromatid
exchanges
in
human
lymphocytes
in
vitro
in
the
absence
of
S9
activation.
Bromodichloromethane
caused
a
dose­
dependent
increase
in
sister
chromatid
exchanges.
The
increased
incidence
was
significant
(
p
<
0.05)
at
concentrations
greater
than
or
equal
to
400
µ
M.
The
potential
of
S9­
activated
bromodichloromethane
to
induce
sister
chromatid
exchanges
in
vitro
was
also
investigated
by
Sobti
(
1984).
A
dose
of
100
µ
M
increased
the
frequency
of
sister
chromatid
exchange
in
rat
liver
cells.
In
human
lymphocytes,
a
100­
fold
greater
concentration
of
bromodichloromethane
was
required
to
elicit
the
same
effect
on
sister
chromatid
exchange
when
compared
to
dibromochloromethane
(
100
µ
M
vs.
1
µ
M.).
Fujie
et
al.
(
1993)
observed
a
statistically
significant,
dose­
related
increase
in
sister
chromatid
exchange
in
rat
erythroblastic
leukemia
K
3
D
cells
treated
with
bromodichloromethane
in
the
absence
of
exogenous
activation.
Bromodichloromethane
also
appeared
to
give
a
positive
response
in
the
presence
of
exogenous
metabolic
activation,
although
the
study
protocol
and
results
with
negative
controls
were
described
less
clearly
for
this
phase
of
testing.

Bromodichloromethane
was
tested
in
the
mouse
lymphoma
assay
as
part
of
an
international
collaborative
program
under
the
auspices
of
the
Japanese
Ministry
of
Health
and
Welfare
(
Sofuni
et
al.,
1996).
The
results
of
this
assay
were
equivocal.
One
laboratory
obtained
a
positive
result
in
the
activated
phase,
but
this
result
was
not
confirmed
by
a
second
laboratory.
Results
in
the
nonactivated
phase
were
negative
or
equivocal
due
to
poor
viability
of
the
solvent
control
cell
cultures.

Matsuoka
et
al.
(
1996)
conducted
a
chromosome
aberration
assay
with
Chinese
hamster
lung
fibroblast
(
CHL/
IU)
cells
exposed
to
bromodichloromethane
in
tightly
capped
flasks.
A
weak
induction
of
chromosome
aberrations
was
observed
for
bromodichloromethane
in
the
presence
and
absence
of
exogenous
metabolic
activation.

Several
studies
have
investigated
the
mutagenicity
of
bromodichloromethane
in
strains
of
Salmonella
typhimurium
engineered
to
express
the
rat
theta­
class
glutathione
S­
transferase
T1­
1
gene
(
GSTT1­
1).
These
studies
provide
evidence
for
a
third
mechanism
of
brominated
trihalomethane
activation
and
thus
are
discussed
in
detail.
Pegram
et
al.
(
1997)
utilized
two
new
strains
of
TA1535­
derived
Salmonella
to
investigate
glutathione
S­
transferase­
mediated
bioactivation
of
bromodichloromethane.
One
strain
had
been
transfected
with
the
GSTT1­
1
gene
(+
GST)
and
the
Draft
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20,
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V
­
71
other
strain
had
the
same
gene
inserted
in
a
non­
functioning
orientation
(­
GST).
These
strains
were
used
in
base­
substitution
revertant
colony
assays
following
24
hour
exposures
to
concentrations
of
bromodichloromethane
ranging
from
200
to
4,800
ppm.
The
agar
concentration
resulting
from
a
24
hour
exposure
to
4,800
ppm
bromodichloromethane
was
0.67
mM.
Bromodichloromethane
increased
the
number
of
revertant
colonies
in
each
strain
of
Salmonella
tested
(+
GST,
­
GST
and
TA1535).
The
frequency
of
the
revertants
in
TA1535
was
significantly
increased
above
the
spontaneous
level
at
the
three
highest
concentrations
tested
(
highest
concentration
4,800
ppm;
intermediate
concentrations
not
explicitly
stated),
while
frequency
was
increased
in
the
­
GST
strain
only
at
the
highest
concentration.
In
contrast,
there
was
a
dramatic,
dose­
dependent
increase
in
bromodichloromethane­
induced
mutations
in
the
+
GST
strain
when
compared
to
the
­
GST
control
strain
(
an
18­
fold
increase
at
the
4800
ppm
bromodichloromethane
concentration).
When
chloroform
was
tested
for
comparative
purposes,
a
positive
response
was
observed
only
at
the
two
highest
concentrations
tested
(
19,200
and
25,600
ppm).
These
results
provide
evidence
that
the
mutagenicity
of
bromodichloromethane
is
enhanced
by
GST­
mediated
conjugation
with
GSH.
The
comparatively
low
affinity
of
the
GSTmediated
pathway
for
chloroform
suggests
that
different
trihalomethanes
can
induce
mutations
by
different
mechanisms.

DeMarini
et
al.
(
1997)
further
investigated
the
role
of
glutathione
S­
transferase
activity
in
mediating
the
mutagenicity
of
bromodichloromethane
in
Salmonella
typhimurium.
Strains
of
Salmonella
utilized
in
this
investigation
included
RSJ100,
which
has
been
engineered
to
express
the
GSTT1­
1
gene
and
TPT100,
which
has
the
GSTT1­
1
gene
inserted
in
a
non­
functioning
orientation.
Mutagenicity
was
assayed
using
a
Tedlar
bag
vaporization
technique.
Bromodichloromethane
(
3,200
ppm)
induced
a
44­
fold
increase
in
revertant
colonies
in
the
RSJ100
strain
of
Salmonella
when
compared
to
background
revertant
frequency.
The
spectrum
of
bromodichloromethane­
induced
mutations
at
the
hisG46
allele
in
strain
RSJ100
was
analyzed
using
the
colony
probe
hybridization
method.
This
analysis
revealed
that
99%
of
the
mutations
were
GC

AT.
A
non­
brominated
halomethane,
dichloromethane,
was
used
in
S.
typhimurium
strain
TA100
(
which
does
not
contain
the
GSST1­
1
expressing
of)
for
comparison.
In
contrast
to
bromodichloromethane­
induced
mutations
in
RSJ100,
only
15%
of
the
mutations
induced
by
dichloromethane
in
the
non­
GST­
expressing
strain
TA100
were
GC

AT
type
mutations.
This
result
suggests
that
over­
expression
of
GSTT1­
1
in
strain
RSJ100
enhanced
the
mutagenicity
of
bromodichloromethane
and
induced
a
specific
type
of
mutational
lesion
in
Salmonella.
The
mutagenicity
of
dibromochloromethane
and
bromoform
was
also
markedly
enhanced
in
the
GSTexpressing
strain
(
discussed
below),
suggesting
that
the
brominated
trihalomethanes
are
bioactivated
by
a
similar
pathway.
In
contrast,
the
mutagenicity
of
chloroform
was
not
enhanced,
indicating
that
chloroform
and
the
brominated
trihalomethanes
may
be
activated
via
different
mechanisms.
Proposed
routes
for
GST­
mediated
bioactivation
are
illustrated
in
Figure
V­
2.
Draft
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or
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February
20,
2003
V
­
72
Figure
V­
2
Proposed
Routes
for
GST­
Mediated
Metabolic
Activation
of
Trihalomethanes
Note:
Solid
arrows
represent
known
pathways
(
modified
from
Gargas
et
al.
(
1986));
dashed
arrows
represent
proposed
pathways
lacking
direct
experimental
evidence.
(
I),
S­(
1­
halomethyl)
GSH;
(
II),
formaldehyde;
(
IIIa),
S­
(
1­
formyl)
GSH;
(
IIIb),
formic
acid;
(
IV),
S­[
1­(
N2­
deoxyguanosinyl)
methyl]
GSH
adduct
(
Thier
et
al.,
1993);
(
V),
S­(
1,1­
dihalomethyl)
GSH;
(
VI),
S­[
1­
halo(
N2­
deoxyguanosinyl)
methyl]
GSH
adduct;
and
(
VII),
N­
formyl
adduct
on
either
G
or
C.
Adapted
from
DeMarini
et
al.
(
1997).

In
Vivo
Assays
Ishidate
et
al.
(
1982)
investigated
the
in
vivo
clastogenicity
of
bromodichloromethane
in
ddY
mice,
MS
mice,
and
Wistar
rats.
Doses
of
125
to
500
mg/
kg­
day
were
administered
in
olive
oil
by
intraperitoneal
injection,
and
the
animals
were
sacrificed
at
18,
24,
30,
48,
and
72
hours
after
dosing.
No
significant
induction
of
micronucleus
formation
in
bone
marrow
was
observed
in
either
mice
or
rats.

Morimoto
and
Koizumi
(
1983)
investigated
the
potential
of
bromodichloromethane
to
induce
sister
chromatid
exchanges
in
male
ICR/
SJ
mice.
Animals
were
given
doses
of
0,
25,
50,
100,
or
200
mg/
kg­
day
for
four
days
by
olive
oil
gavage.
Bromodichloromethane
produced
a
roughly
linear
dose­
dependent
increase
in
sister
chromatid
exchange
frequency.
The
increase
was
statistically
significant
(
p
<
0.05)
at
50
mg/
kg­
day.
The
authors
noted
that
the
concentrations
required
to
produce
an
increased
incidence
of
sister
chromatid
exchange
were
on
the
order
of
1,000
to
10,000
times
higher
than
the
concentrations
typically
found
in
drinking
water.

Hayashi
et
al.
(
1988)
measured
induction
of
micronucleated
polychromatic
erythrocytes
in
ddY
mice
by
intraperitoneal
administration
of
bromodichloromethane
at
single
doses
up
to
500
Draft
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20,
2003
V
­
73
mg/
kg
in
corn
oil.
No
evidence
of
clastogenicity
was
observed.
There
was
no
clear
evidence
of
toxicity
or
cytotoxicity
in
the
target
tissue.

Fujie
et
al.
(
1990)
analyzed
chromosome
aberrations
in
bone
marrow
from
Long­
Evans
rats
(
3/
sex/
dose)
following
oral
(
males
only)
or
intraperitoneal
(
males
and
females)
exposure
to
bromodichloromethane.
Oral
administration
was
by
gavage
in
saline
for
five
consecutive
days,
and
the
animals
were
sacrificed
18
hours
after
the
last
dose.
Bromodichloromethane
induced
dose­
related
increases
in
chromatid
and
chromosome
breaks.
A
more
pronounced
increase
in
clastogenic
activity
was
observed
following
a
single
intraperitoneal
dose,
with
significant
(
p
<

0.05)
effects
at
16.4
mg/
kg.

Hayashi
et
al.
(
1992)
evaluated
induction
of
micronuclei
in
mouse
peripheral
blood
erythrocytes
by
bromodichloromethane.
Groups
of
four
male
ddY
mice
received
an
intraperitoneal
injection
of
0,
25,
50,
or
100
mg/
kg
bromodichloromethane
in
physiological
saline
once
a
week
for
5
weeks.
Micronuclei
were
evaluated
1
week
after
the
last
dose.
No
evidence
of
micronucleus
induction
was
observed.
One
low­
dose
mouse
died,
and
weight
loss
was
observed
in
all
treatment
groups
during
exposure.

Potter
et
al.
(
1996)
investigated
the
effect
of
bromodichloromethane
on
DNA
strand
breakage
in
male
F344
rats.
Test
animals
received
0.75
or
1.5
mmol/
kg
of
bromodichloromethane
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days.
The
administered
doses
corresponded
to
123
or
246
mg/
kg­
day.
One
day
after
administration
of
a
single
dose,
DNA
strand
breaks
in
the
kidney
were
analyzed
using
the
alkaline
unwinding
procedure.
No
treatment­
related
effect
was
observed
at
either
dose
level.

Stocker
et
al.
(
1997)
investigated
the
in
vivo
genotoxicity
of
bromodichloromethane
in
an
unscheduled
DNA
synthesis
assay
in
the
livers
of
bromodichloromethane
treated
rats.
Male
Sprague­
Dawley
rats
(
4
animals
per
group)
were
administered
a
single
dose
of
0
(
control),
135
or
450
mg/
kg
bromodichloromethane
via
gavage
in
aqueous
1%
methylcellulose.
These
doses
were
selected
by
the
authors
to
correspond
to
30%
and
100%
of
the
calculated
maximum
tolerated
dose
(
MTD)
for
this
compound.
Analysis
of
hepatocytes
for
unscheduled
DNA
synthesis
was
conducted
2
and
14
hours
after
treatment.
There
was
no
evidence
of
increased
DNA
synthesis
in
hepatocytes
at
any
tested
dose
of
bromodichloromethane.
Draft
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2003
V
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Table
V­
10
Summary
of
Mutagenicity,
Genotoxicity,
and
Neoplastic
Transformation
Data
for
Bromodichloromethane
Endpoint
Assay
System
Results
(
with/
without
activation)
d
References
In
Vitro
Studies
Gene
mutation
Salmonella
typhimurium
TA100a
NT/+
Simmon
and
Tardiff
(
1978)

TA100b
­/+
Ishidate
et
al.
(
1982)

TA98,
TA100,
TA1535,
TA1537b
­/­
NTP
(
1987)

TA1537
TA1535,
TA98,
TA100b
­/+
­/­
Varma
et
al.
(
1988)

RSJ100
NT/+
DeMarini
et
al.
(
1997)

TA1535,
+
GST,
­
GST
NT/+
Pegram
et
al.
(
1997)

Mouse
lymphoma
cellsb
+/­
NTP
(
1987)

Mouse
lymphoma
cells
+
c/­
c
Sofuni
et
al.
(
1996)

Chromosome
aberration
Chinese
hamster
fibroblasts
b
+/­
Ishidate
et
al.
(
1982)

Chinese
hamster
ovary
cells
b
­/­
NTP
(
1987);
Anderson
et
al.
(
1990)

Chinese
hamster
lung
fibroblasts
a
+/+
(
weak)
Matsuoka
et
al.
(
1996)

DNA
damage
Saccharomyces
cerevisiae
a
­/­
Nestmann
and
Lee
(
1985)

SOS
chromotest
+/+
LeCurieux
et
al.
(
1995)

Sister
chromatid
exchange
Human
lymphocytes
a
NT/+
Morimoto
and
Koizumi
(
1983)

Human
lymphocytes
a
+/
NT
Sobti
(
1984)

Rat
liver
cells
a
+/
NT
Sobti
(
1984)

Chinese
hamster
ovary
cells
b
­
c/­
NTP
(
1987);
Anderson
et
al.
(
1990)

In
Vivo
Studies
Micronuclei
Mouse
bone
marrow
cells
and
rat
cells
­
Ishidate
et
al.
(
1982)

Mouse
bone
marrow
cells
­
Hayashi
et
al.
(
1988)
Table
V­
10
(
cont.)

Endpoint
Assay
System
Results
(
with/
without
activation)
d
References
Draft
­
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Not
Cite
or
Quote
February
20,
2003
V
­
75
Mouse
peripheral
blood
erythrocytes
(
ip)
­
Hayashi
et
al.
(
1992)

Chromosome
aberrations
Rat
bone
marrow
cells
(
oral)
+
Fujie
et
al.
(
1990)

Rat
bone
marrow
cells
(
ip)
+
Fujie
et
al.
(
1990)

Rat
kidney
cells
­
Potter
et
al.
(
1996)

Unscheduled
DNA
synthesis
Rat
liver
cells
­
Stocker
et
al.
(
1997)

Sister
chromatid
exchange
Mouse
bone
marrow
cells
+
Morimoto
and
Koizumi
(
1983)

NT
=
Not
Tested
a
Assay
was
conducted
in
a
closed
system.
b
Authors
did
not
specify
whether
or
not
the
assay
was
conducted
in
a
closed
system.

c
Equivocal
results
were
obtained.
d
With/
without
activation
applies
to
in
vitro
tests
only.
Draft
­
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or
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20,
2003
V
­
76
2.
Dibromochloromethane
The
results
of
in
vivo
and
in
vitro
tests
conducted
to
evaluate
the
mutagenicity,
genotoxicity,
and
neoplastic
transformation
potential
of
dibromochloromethane
are
summarized
in
Table
V­
11
at
the
end
of
this
section.

In
Vitro
Assays
Simmon
and
Tardiff
(
1978)
reported
that
nonactivated
dibromochloromethane
was
mutagenic
in
S.
typhimurium
strain
TA100
when
assayed
in
a
desiccator
containing
the
test
compound
in
the
atmosphere.
The
minimum
amount
of
dibromochloromethane
required
to
elicit
a
mutagenic
response
following
addition
to
the
desiccator
was
57
µ
mol.

Ishidate
et
al.
(
1982)
assayed
the
mutagenicity
of
dibromochloromethane
in
S.
typhimurium
strain
TA100
in
the
presence
and
absence
of
rat
liver
S9
fraction.
Increased
mutation
frequencies
were
observed
only
in
the
absence
of
S9
activation.
In
contrast,
chromosomal
aberrations
in
Chinese
hamster
fibroblasts
were
observed
in
the
presence,
but
not
the
absence,
of
S9
fraction.
The
concentrations
tested
in
these
assays
were
not
reported.

NTP
(
1985)
reported
that
dibromochloromethane
(
0.5
to
50
µ
mol/
plate;
100
to
10,000
µ
g/
plate)
was
not
mutagenic
in
strains
TA1535,
TA1537,
TA98,
or
TA100
when
tested
in
the
presence
or
absence
of
Aroclor­
induced
Sprague­
Dawley
rat
or
Syrian
hamster
liver
S9
fractions.
Volatilization
of
the
test
compound
was
proposed
as
a
possible
explanation
for
the
negative
results.

Nestmann
and
Lee
(
1985)
investigated
the
mutagenicity
of
dibromochloromethane
at
concentrations
of
11
to
5,700
µ
M
in
S.
cerevisiae
strains
D7
and
XV185­
14C
in
the
presence
or
absence
of
S9
activation.
No
clear
increase
in
convertants
or
in
revertants
of
strain
XV185­
14C
were
observed
in
the
presence
of
S9­
activated
dibromochloromethane.
In
the
absence
of
S9
activation,
an
increased
incidence
of
gene
convertants
in
strain
D7
was
observed
at
concentrations
greater
than
1,140
µ
M.
There
was
no
effect
on
the
incidence
of
revertants
under
the
same
conditions.
The
high
dose
of
dibromochloromethane
was
cytotoxic.

Varma
et
al.
(
1988)
tested
dibromochloromethane
for
mutagenicity
in
S.
typhimurium
strains
TA1535,
TA1537,
TA98,
and
TA100.
Dibromochloromethane
produced
a
significantly
increased
mutation
frequency
at
the
lowest
S9­
activated
concentration
(
0.12
µ
mol/
plate)
in
all
four
strains.
Dibromochloromethane
at
the
same
concentration
also
resulted
in
increased
mutation
frequencies
in
strains
TA1535
and
TA1537
in
the
absence
of
S9
fraction.
Higher
concentrations
had
no
clear
effect
on
mutation
frequency.
This
spike
in
mutation
frequency
at
the
low
dose
with
similar
responses
in
strains
that
detect
frameshifts
and
those
that
detect
base
substitutions
is
very
unusual.
It
is
possible
that
the
reported
data
may
have
resulted
from
cytotoxicity,
although
the
number
of
revertants
at
the
nonmutagenic
doses
was
comparable
to
background
levels.
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V
­
77
Dibromochloromethane
induced
mutations
at
the
tk
locus
of
L5178Y
mouse
lymphoma
cells
when
tested
at
concentrations
greater
than
or
equal
to
480
µ
M
in
screw­
capped
tubes.
The
material
was
tested
only
in
the
absence
of
S9
activation
(
McGregor
et
al.,
1991).

Loveday
et
al.
(
1990)
found
that
dibromochloromethane
did
not
induce
chromosome
aberrations
in
CHO
cells
with
S9­
activation
at
concentrations
that
caused
cell­
cycle
delay
(
12,200
µ
M)
or
in
the
absence
of
S9­
activation
at
concentrations
that
were
cytotoxic
with
a
standard
harvest
time
(
6,000
µ
M).
Sister
chromatid
exchange
was
induced
in
CHO
cells
by
S9­
activated
dibromochloromethane
at
3,600
µ
M
with
a
delayed
cell
harvest,
while
the
nonactivated
test
material
had
no
effect
at
concentrations
up
to
cytotoxic
levels
(
1,200
µ
M;
247
µ
g/
mL).

Morimoto
and
Koizumi
(
1983)
investigated
the
ability
of
dibromochloromethane
to
induce
sister
chromatid
exchanges
(
SCE)
in
human
lymphocytes
in
vitro
in
the
absence
of
S9
activation.
Addition
of
dibromochloromethane
resulted
in
a
dose­
dependent
increase
in
SCE.
The
increased
incidence
was
significant
(
p
<
0.05)
at
concentrations
greater
than
or
equal
to
400
µ
M.

The
potential
of
S9­
activated
dibromochloromethane
to
induce
sister
chromatid
exchanges
in
vitro
was
also
investigated
by
Sobti
(
1984).
A
dose
of
100
µ
M
produced
an
increased
frequency
of
sister
chromatid
exchange
in
rat
liver
cells.
In
human
lymphocytes,
1
µ
M
dibromochloromethane
produced
the
same
effect
as
100
µ
M
bromodichloromethane.

Fujie
et
al.
(
1993)
observed
a
statistically
significant,
dose­
related
increase
in
sister
chromatid
exchange
in
rat
erythroblastic
leukemia
K
3
D
cells
treated
with
dibromochloromethane
in
the
absence
of
exogenous
activation.
Dibromochloromethane
had
the
weakest
response
among
the
brominated
trihalomethanes
tested.
Dibromochloromethane
also
appeared
to
give
a
positive
response
in
the
presence
of
exogenous
metabolic
activation,
although
the
study
protocol
and
results
with
negative
controls
were
less
clear
for
this
phase
of
testing.

LeCurieux
et
al.
(
1995)
evaluated
the
induction
of
DNA
damage
by
dibromochloromethane
in
the
presence
and
absence
of
exogenous
activation
using
the
SOS
chromotest.
Dibromochloromethane
exposure
gave
a
positive
result
for
induction.
Dibromochloromethane
gave
negative
results
in
the
fluctuation
test
modification
of
the
S.
typhimurium
reverse
mutation
assay.

Matsuoka
et
al.
(
1996)
conducted
a
chromosome
aberration
assay
with
Chinese
hamster
lung
fibroblast
(
CHL/
IU)
cells
exposed
to
dibromochloromethane
in
tightly
capped
flasks.
Dibromochloromethane
induced
polyploidy
in
the
absence
of
S9
fraction,
but
not
in
the
presence
of
S9.
The
study
authors
considered
activated
dibromochloromethane
marginally
positive
for
chromosome
aberrations.
However,
there
was
no
effect
under
the
utilized
test
conditions
when
gaps
were
excluded
from
consideration.
There
was
no
evidence
of
structural
chromosome
aberration
induction
by
dibromochloromethane
in
the
absence
of
exogenous
metabolic
activation.

Dibromochloromethane
was
tested
in
the
mouse
lymphoma
assay
as
part
of
an
international
collaborative
program
under
the
auspices
of
the
Japanese
Ministry
of
Health
and
Draft
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or
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February
20,
2003
V
­
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Welfare
(
Sofuni
et
al.,
1996).
Dibromochloromethane
yielded
clearly
positive
results
with
or
without
exogenous
metabolic
activation
in
two
laboratories.

DeMarini
et
al.
(
1997)
investigated
the
role
of
glutathione
S­
transferase
activity
in
the
mutagenicity
of
dibromochloromethane
in
Salmonella
typhimurium.
Strains
of
Salmonella
utilized
in
this
investigation
included
RSJ100,
which
is
engineered
to
express
the
rat
glutathione
S­
transferase
theta
1­
1
(
GSTT1­
1)
gene
and
TPT100,
which
has
the
GSTT1­
1
gene
inserted
in
a
non­
functioning
orientation.
Dibromochloromethane
(
400
ppm)
induced
an
85­
fold
increase
in
revertant
colonies
in
the
RSJ100
strain
of
Salmonella
compared
to
background
revertant
formation.
The
mutational
spectra
for
dibromochloromethane­
induced
mutations
at
the
hisG46
allele
in
strain
RSJ100
were
analyzed
using
the
colony
probe
hybridization
method.
This
analysis
revealed
that
100%
of
the
mutations
were
GC

AT.
A
non­
brominated
dihalomethane,
dichloromethane,
was
tested
in
TA100
(
which
does
not
express
GSTT1­
1)
for
comparison.
In
contrast
to
dibromochloromethane­
induced
mutations
in
RSJ100,
only
15%
of
the
mutations
induced
by
dichloromethane
in
TA100
were
GC

AT
type
mutations.
This
result
suggests
that
over­
expression
of
GSTT1­
1
in
strain
RSJ100
mediated
the
mutagenicity
of
dibromochloromethane
and
induced
a
specific
type
of
mutational
lesion
in
Salmonella.
Proposed
pathways
of
bioactivation
of
dibromochloromethane
and
other
brominated
trihalomethanes
are
shown
in
Figure
4­
2.

Landi
et
al.
(
1999)
investigated
the
role
of
GSST1­
1
in
the
mutagenicity
of
dibromochloromethane
in
Salmonella
by
using
one
strain
that
expressed
rat
GSST1­
1
(
RSJ100)
and
one
strain
that
did
not
(
TPT100).
Mutagenicity
of
dibromochloromethane
was
assessed
by
revertant
colony
formation
with
or
without
S9
metabolic
activation.
The
addition
of
800
ppm
dibromochloromethane
greatly
increased
revertant
numbers
in
the
RSJ100
but
not
the
TPT100
strain
of
Salmonella.
Addition
of
the
rat
liver
S9
fraction
had
no
effect
on
the
number
of
revertants
induced
by
dibromochloromethane
exposure
in
either
strain.
These
data
provide
further
support
for
the
hypothesis
that
GSST1­
1
plays
a
role
in
the
mutagenicity
of
dibromochloromethane.
Additional
experiments
were
conducted
to
investigate
the
effects
of
exogenously
added
GSST1­
1
on
the
mutagenic
potency
of
dibromochloromethane.
Red
blood
cells
(
RBC),
which
express
GSST1­
1,
were
added
to
the
experimental
system
to
address
this
question.
RBC
had
no
effect
on
results
obtained
with
the
TPT100
strain,
but
completely
suppressed
the
mutagenicity
of
dibromochloromethane
in
the
RSJ100
strain.
However,
the
>

protective=
effect
of
RBC
did
not
appear
to
be
related
to
GSST1­
1
activity,
as
this
suppression
occurred
even
with
the
addition
of
RBC
from
individuals
who
do
not
express
GSST1­
1.
The
underlying
mechanism
of
RBC
suppression
of
dibromochloromethane
mutagenicity
was
not
investigated.
The
authors
of
this
study
speculated
that
tissues
potentially
exposed
to
dibromochloromethane
via
the
blood
may
be
at
less
genotoxic
risk
(
due
to
protection
afforded
by
the
RBC)
than
tissues
which
are
directly
exposed
to
oral
bromodichloromethane
(
such
as
tissues
in
the
gastrointestinal
tract).
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February
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2003
V
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79
In
Vivo
Assays
Fujie
et
al.
(
1990)
analyzed
chromosome
aberrations
in
bone
marrow
from
Long­
Evans
rats
(
3/
sex/
dose)
following
oral
(
males
only)
or
intraperitoneal
(
males
and
females)
exposure
to
dibromochloromethane.
Oral
administration
was
by
gavage
in
saline
for
five
consecutive
days,
and
the
animals
were
sacrificed
18
hours
after
the
last
dose.
Dibromochloromethane
induced
dose­
related
increases
in
chromosome
breaks.
A
more
pronounced
increase
in
clastogenic
activity
was
observed
following
a
single
intraperitoneal
dose,
with
significant
(
p
<
0.05)
effects
at
20.8
mg/
kg.
Regardless
of
the
route,
the
predominant
types
of
induced
aberrations
were
chromatid
and
chromosome
breaks.

Hayashi
et
al.
(
1988)
measured
induction
of
micronucleated
polychromatic
erythrocytes
in
ddY
mice
by
intraperitoneal
administration
of
dibromochloromethane
at
single
doses
of
up
to
500
mg/
kg
in
corn
oil.
No
evidence
of
clastogenicity
was
observed.
However,
the
sampling
time
utilized
in
this
experiment
was
insufficient
(
U.
S.
EPA,
1994b).
There
was
no
clear
evidence
of
toxicity
or
cytotoxicity
in
the
target
tissue.

Ishidate
et
al.
(
1982)
investigated
the
in
vivo
clastogenicity
of
dibromochloromethane
in
ddY
and
MS
mice
and
Wistar
rats.
Doses
of
125
to
500
mg/
kg­
day
were
administered
in
olive
oil
by
intraperitoneal
injection,
and
the
animals
were
sacrificed
at
18,
24,
30,
48,
and
72
hours
after
dosing.
No
significant
induction
of
micronucleus
formation
was
observed
in
either
mice
or
rats.

Morimoto
and
Koizumi
(
1983)
investigated
the
potential
of
dibromochloromethane
to
induce
sister
chromatid
exchanges
in
male
ICR/
SJ
mice.
Animals
were
given
doses
of
0,
25,
50,
100,
or
200
mg/
kg­
day
for
four
days
by
olive
oil
gavage.
Dibromochloromethane
produced
a
roughly
linear
dose­
dependent
increase
in
sister
chromatid
exchange
frequency.
The
increase
was
statistically
significant
(
p
<
0.05)
at
25
mg/
kg­
day.
The
authors
noted
that
the
concentrations
required
to
produce
an
increased
incidence
of
sister
chromatid
exchange
were
on
the
order
of
1,000
to
10,000
times
higher
than
the
concentrations
typically
found
in
drinking
water.

Potter
et
al.
(
1996)
dosed
male
F344
rats
(
4/
dose)
with
0.75
or
1.5
mmol/
kg
of
dibromochloromethane
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days
and
investigated
several
endpoints
potentially
related
to
kidney
tumorigenesis.
These
doses
corresponded
to
156
or
312
mg/
kg­
day.
No
effect
was
observed
when
DNA
strand
breaks
in
the
kidney
were
analyzed
using
the
alkaline
unwinding
procedure
one
day
following
treatment
with
a
single
dose
of
dibromochloromethane.
Because
kidney
tumors
induced
by
some
chemicals
in
male
rats
have
been
related
to
the
formation
of
 2u­
globulin
rich
hyaline
droplets,
kidney
hyaline
droplets
were
also
evaluated
in
all
of
the
dosed
rats.
Binding
to
 2u­
globulin
was
not
measured.
No
exposurerelated
increase
in
hyaline
droplets
was
found.
Changes
in
kidney
tubule
cell
proliferation
were
assessed
by
in
vivo
incorporation
[
3H]­
thymidine.
No
statistically
significant
effect
of
dibromochloromethane
exposure
on
this
endpoint
following
was
noted
exposures
of
up
to
7
days
duration.
Draft
­
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February
20,
2003
V
­
80
Stocker
et
al.
(
1997)
investigated
the
in
vivo
genotoxicity
of
dibromochloromethane
in
an
unscheduled
DNA
synthesis
assay
in
the
livers
of
dibromochloromethane
treated
rats.
Male
Sprague­
Dawley
rats
(
4
animals
per
group)
were
administered
a
single
dose
of
0
(
control),
600
or
2000
mg/
kg
via
gavage
in
aqueous
1%
methylcellulose.
These
doses
were
selected
by
the
authors
to
correspond
to
30%
and
100%
of
the
calculated
maximum
tolerated
dose
(
MTD)
for
this
compound.
Analysis
of
hepatocytes
for
unscheduled
DNA
synthesis
was
conducted
2
and
14
hours
after
treatment.
There
was
no
evidence
of
increased
DNA
synthesis
in
hepatocytes
from
rats
treated
with
any
tested
dose
of
dibromochloromethane.

Sekihashi
et
al.
(
2002)
obtained
positive
results
for
dibromochloromethane
genotoxicity
using
the
Comet
assay.
The
authors
indicated
that
doses
were
selected
to
avoid
confounding
of
the
results
by
cytotoxicity.
In
Wistar
rats,
positive
(
statistically
significant
differences
in
mean
migration)
results
were
obtained
for
stomach,
colon,
liver,
kidney,
bladder,
or
lung
tissues
removed
8
or
24
hours
following
administration
of
200
mg/
kg
oral
dose
of.
using.
In
ddY
mice,
positive
results
were
obtained
for
liver
and
brain
samples
harvested
8
or
24
hours,
respectively,
after
administration
of
a
400
mg/
kg
oral
dose.
Although
a
statistically
significant
increase
in
migration
was
also
noted
for
the
eight
hour
colon
sample,
the
study
authors
did
not
identify
this
finding
as
a
positive
response.
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February
20,
2003
V
­
81
Table
V­
11
Summary
of
Mutagenicity,
Genotoxicity,
and
Neoplastic
Transformation
Data
for
Dibromochloromethane
Endpoint
Assay
System
Results
(
with/
without
activation)
d
References
In
Vitro
Studies
Gene
mutation
Salmonella
typhimurium
TA100a
NT/+
Simmon
and
Tardiff
(
1978)

TA100b
­/+
Ishidate
et
al.
(
1982)

TA98,
TA100,
TA1535,
TA1537b
­/­
NTP
(
1985)

TA1535,
TA1537b
TA98,
TA100b
+/+
­/+
Varma
et
al.
(
1988)

RSJ100
NT/+
DeMarini
et
al.
(
1997)

RSJ100
TPT100
+/+
­/­
Landi
et
al.
(
1999)

Mouse
lymphoma
cellsa
NT/+
McGregor
et
al.
(
1991)

Mouse
lymphoma
cells
+/+
Sofuni
et
al.
(
1996)

Chromosome
aberration
Chinese
hamster
fibroblastsb
+/­
Ishidate
et
al.
(
1982)

Chinese
hamster
ovary
cellsb
­/­
Loveday
et
al.
(
1990)

Chinese
hamster
lung
fibroblastsa
­/+
(
see
text)
Matsuoka
et
al.
(
1996)

DNA
damage
Saccharomyces
cerevisiaea
­/+
Nestmann
and
Lee
(
1985)

SOS
chromotest
S.
typhimurium
fluctuation
test
+/+
­
LeCurieux
et
al.
(
1995)

Sister
chromatid
exchange
Human
lymphocytesa
NT/+
Morimoto
and
Koizumi
(
1983)

Human
lymphocytesa
+/
NT
Sobti
(
1984)

Rat
liver
cellsb
+/
NT
Sobti
(
1984)

Chinese
hamster
ovary
cellsb
+/­
Loveday
et
al.
(
1990)

Rat
erythroblastic
leukemia
cells
­
c/+
Fujie
et
al.
(
1993)

In
Vivo
Studies
Micronuclei
Mouse
bone
marrow
cells
­
Ishidate
et
al.
(
1982)
Table
V­
11
(
cont.)

Endpoint
Assay
System
Results
(
with/
without
activation)
d
References
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
82
Mouse
bone
marrow
cells
­
Hayashi
et
al.
(
1988)

Chromosome
aberrations
Rat
bone
marrow
cells
+
Fujie
et
al.
(
1990)

Rat
bone
marrow
cells
+
Fujie
et
al.
(
1990)

Sister
chromatid
exchange
Mouse
bone
marrow
cells
+
Morimoto
and
Koizumi
(
1983)

DNA
damage
Rat
kidney
cells
­
Potter
et
al.
(
1996)

Rat
stomach,
colon,
liver,
kidney,
bladder,
lung
tissue
+
Sekihashi
et
al.
(
2002)

Mouse
liver
and
brain
tissue
+
Sekihashi
et
al.
(
2002)

Unscheduled
DNA
synthesis
Rat
hepatocytes
­
Stocker
et
al.
(
1997)

NT
=
Not
Tested
a
Assay
was
conducted
in
a
closed
system.
b
Authors
did
not
specify
whether
or
not
the
assay
was
conducted
in
a
closed
system.
c
Equivocal
results
reported.
d
With/
without
activation
applies
to
in
vitro
data
only.
Draft
­
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February
20,
2003
V
­
83
3.
Bromoform
The
results
of
in
vivo
and
in
vitro
tests
conducted
to
evaluate
the
mutagenicity,
genotoxicity,
and
neoplastic
transformation
potential
of
bromoform
are
summarized
in
Table
V­
12
at
the
end
of
this
section.

In
Vitro
Assays
Simmon
and
Tardiff
(
1978)
reported
that
nonactivated
bromoform
was
mutagenic
in
S.
typhimurium
strain
TA100
when
assayed
as
vapor
in
a
desiccator.
The
minimum
amount
of
bromoform
required
to
elicit
a
mutagenic
response
following
addition
to
the
desiccator
was
570
µ
mol.

Ishidate
et
al.
(
1982)
assayed
the
mutagenicity
of
bromoform
in
S.
typhimurium
strain
TA100
in
the
presence
and
absence
of
rat
liver
S9
fraction.
Increased
mutation
frequencies
were
observed
only
in
the
absence
of
S9
activation.
In
contrast,
chromosomal
aberrations
in
Chinese
hamster
fibroblasts
were
observed
in
the
presence,
but
not
the
absence,
of
S9
fraction.
The
concentrations
tested
in
these
assays
were
not
reported.

Maddock
and
Kelly
(
1980)
reported
that
bromoform
did
not
induce
an
increase
in
sister
chromatid
exchanges
when
toadfish
leukocytes
were
exposed
to
concentrations
of
0.4
to
400
µ
M.

Herren­
Freund
and
Pereira
(
1986)
assessed
the
initiating
activity
of
bromoform
using
the
rat
liver
GGT­
foci
assay.
The
authors
reported
that
a
250
mg/
kg
(
1
mmol/
kg)
oral
dose
in
an
unspecified
vehicle
did
not
initiate
GGT­
foci
in
this
test.

NTP
(
1989a)
evaluated
the
genotoxic
potential
of
bromoform
in
multiple
test
systems.
Concentrations
of
0.04
to
13
µ
mol/
plate
(
10
to
3,333
µ
g/
plate)
produced
no
evidence
of
mutagenicity
in
S.
typhimurium
strains
TA1535
or
TA1537,
when
assayed
with
or
without
exogenous
metabolic
activation
by
rat
or
hamster
liver
S9
fraction.
Equivocal
evidence
of
mutagenicity
was
noted
in
strain
TA100
without
activation,
and
in
strains
TA97
and
TA98
in
the
presence
of
liver
microsomes
prepared
from
Aroclor­
induced
Syrian
hamsters.
Exposure
of
mouse
L5178Y
cells
to
bromoform
concentrations
greater
than
or
equal
to
2,300
µ
M
in
the
absence
of
S9
activation
or
S9­
activated
concentrations
of
at
least
300
µ
M
with
S9
activation
resulted
in
forward
mutations
at
the
thymidine
kinase
(
tk)
locus.
One
of
two
laboratories
conducting
the
assays
reported
increased
sister
chromatid
exchanges
(
SCE)
in
CHO
cells
exposed
to
3,800
µ
M
bromoform
in
the
absence
of
exogenous
activation.
Neither
laboratory
observed
increased
incidence
of
SCE
in
the
presence
of
S9.
S9­
activated
bromoform
did
not
induce
chromosome
aberrations
in
CHO
cells;
results
for
SCE
and
chromosome
aberrations
in
the
absence
of
exogenous
activation
were
equivocal.

Zeiger
(
1990)
found
that
bromoform
was
mutagenic
in
S.
typhimurium
strain
TA98
when
tested
as
a
vapor
in
a
closed
system,
but
not
when
tested
in
an
open
system
using
a
preincubation
protocol.
Positive
results
were
observed
at
levels
of
at
least
114
µ
mol/
desiccator,
in
the
presence
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February
20,
2003
V
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84
and
absence
of
S9
prepared
from
rat
or
hamster
liver.
Bromoform
was
negative
in
the
closed
system
with
strains
TA100
and
TA1538
with
or
without
rat
or
hamster
liver
S9
fraction
Roldan­
Arjona
and
Pueyo
(
1993)
evaluated
bromoform
in
the
S.
typhimurium
Ara
forward
mutation
assay
at
concentrations
up
to
25
µ
mol/
plate
(
6.3
mg/
plate).
A
preincubation
protocol
was
employed
for
the
assay.
Although
a
clear
dose­
related
response
was
observed
in
the
absence
of
activation,
the
results
were
classified
as
questionable
because
a
doubling
of
background
levels
was
not
achieved.
There
was
no
evidence
of
mutagenicity
in
the
presence
of
exogenous
metabolic
activation.
Although
no
attempt
was
made
to
minimize
volatilization
of
the
test
compound,
cytotoxicity
at
the
high
exposure
level
indicated
that
the
test
material
reached
the
cells.

DeMarini
et
al.
(
1997)
investigated
the
role
of
glutathione
S­
transferase
activity
in
the
mutagenicity
of
bromoform
in
Salmonella
typhimurium.
Strains
of
Salmonella
utilized
for
this
investigation
included
RSJ100,
which
has
been
engineered
to
express
the
rat
glutathione
Stransferase
theta
1­
1
(
GSTT1­
1)
gene
and
TPT100,
which
has
the
GSTT1­
1
gene
inserted
in
a
non­
functioning
orientation.
Exposure
to1,600
ppm
bromoform
induced
a
95­
fold
increase
in
revertant
colonies
in
the
RSJ100
strain
of
Salmonella
compared
to
background
revertant
formation.
The
mutational
spectra
for
bromoform­
induced
mutations
at
the
hisG46
allele
in
strain
RSJ100
were
analyzed
using
the
colony
probe
hybridization
method.
This
analysis
revealed
that
96%
of
the
mutations
were
GC

AT
transitions.
Bromoform
also
induced
a
smaller
percentage
(
2.8%)
of
GC

TA
mutations.
A
non­
brominated
halomethane,
dichloromethane,
was
used
in
S.
typhimurium
strain
TA100
(
which
does
not
express
GSST1­
1)
for
comparison.
In
contrast
to
bromoform­
induced
mutations
in
RSJ100,
only
15%
of
the
mutations
induced
by
dichloromethane
in
TA100
were
GC

AT
type
mutations.
This
result
suggests
that
over­
expression
of
GSTT1­
1
in
strain
RSJ100
mediated
the
mutagenicity
of
bromoform
and
induced
a
specific
type
of
mutational
lesion
in
Salmonella.
Proposed
pathways
for
the
bioactivation
of
bromoform
and
other
brominated
trihalomethanes
are
illustrated
in
Figure
V­
2.

Landi
et
al.
(
1999)
investigated
the
mutagenicity
of
bromoform
in
in
vitro
exposed
human
lymphocytes
from
both
glutathione­
S­
transferase
theta
positive
(
GSST1­
1+)
and
negative
(
GSST1­
1­)
individuals.
Whole
blood
cultures
were
exposed
to
bromoform
(
10­
2
to
10­
4
M)
and
assayed
for
DNA
breaks
with
the
COMET
assay.
The
DNA­
damaging
potency
of
bromoform
was
not
significantly
different
in
lymphocytes
(
the
target
cell
for
the
COMET
assay)
from
GSST1­
1+
and
GSST1­
1­
individuals.
However,
lymphocytes
do
not
express
GSST1­
1,
even
in
GSST1­
1+
individuals,
so
interpretation
of
this
data
is
problematic.
When
data
were
combined
from
both
genotypic
groups,
there
was
a
weak
but
statistically
significant
induction
of
comets
observed
following
treatment
with
bromoform.

In
Vivo
Assays
NTP
(
1989a)
studied
the
genotoxic
potential
of
bromoform
in
several
test
systems.
Adult
male
Drosophila
fed
with
a
1,000­
ppm
solution
of
bromoform
exhibited
increased
frequency
of
sex­
linked
recessive
lethal
mutations,
but
no
significant
effect
on
reciprocal
translocations
was
Draft
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February
20,
2003
V
­
85
observed.
Intraperitoneal
injection
of
mice
with
200
to
800
mg/
kg
bromoform
caused
an
increase
in
sister
chromatid
exchange
but
not
in
chromosomal
aberrations
in
bone
marrow
cells.

Fujie
et
al.
(
1990)
analyzed
chromosome
aberrations
in
bone
marrow
from
Long­
Evans
rats
(
3/
sex/
dose)
following
oral
(
males
only)
or
intraperitoneal
(
males
and
females)
exposure
to
bromoform.
Oral
administration
was
by
gavage
in
saline
for
five
consecutive
days,
and
the
animals
were
sacrificed
18
hours
after
the
last
dose.
Bromoform
induced
a
dose­
related
increase
in
the
incidence
of
aberrant
cells,
with
a
significant
(
p
<
0.01)
increase
at
253
mg/
kg­
day.
A
more
pronounced
increase
in
clastogenic
activity
was
observed
following
a
single
intraperitoneal
dose,
with
a
significant
(
p
<
0.05)
effect
at
25.3
mg/
kg.
Regardless
of
the
route,
the
predominant
types
of
induced
aberrations
were
chromatid
and
chromosome
breaks.

Morimoto
and
Koizumi
(
1983)
investigated
the
ability
of
bromoform
and
other
brominated
trihalomethanes
to
induce
sister
chromatid
exchanges
in
human
lymphocytes
in
vitro
in
the
absence
of
S9
activation.
All
three
brominated
trihalomethanes
caused
a
dose­
dependent
increase
in
sister
chromatid
exchanges.
Bromoform
was
more
potent
than
bromodichloromethane
or
dibromochloromethane.
The
increases
were
significant
(
p
<
0.05)
at
concentrations
greater
than
or
equal
to
400
µ
M,
400
µ
M,
and
80
µ
M
for
bromodichloromethane,
dibromochloromethane
and
bromoform,
respectively.

Potter
et
al.
(
1996)
evaluated
the
effect
of
bromoform
on
incidence
of
DNA
strand
breaks
in
the
kidney.
Male
F344
rats
received
0.75
or
1.5
mmol/
kg
of
bromoform
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days.
These
doses
corresponded
to
190
or
379
mg/
kg.
No
effect
was
observed
on
strand
breaks
when
evaluated
using
the
alkaline
unwinding
procedure
one
day
after
a
single
dose.

Stocker
et
al.
(
1997)
investigated
the
in
vivo
genotoxicity
of
bromoform
in
the
mouse
bone
marrow
micronuclei
assay
and
by
analysis
of
unscheduled
DNA
synthesis
in
the
liver
of
bromoform­
treated
rats.
In
the
first
assay,
Swiss
CD
mice
(
5/
sex/
dose)
were
treated
by
gavage
with
doses
of
0,
250,
500,
or
1,000
mg/
kg
bromoform
dissolved
in
aqueous
1%
methylcellulose.
Micronuclei
analysis
was
conducted
24
and
48
hours
after
dosing,
and
was
negative
in
all
dose
groups.
In
the
second
assay,
male
Sprague­
Dawley
rats
(
4
animals/
dose)
received
single
doses
of
0,
324
or
1,080
mg/
kg
bromoform
by
gavage
in
aqueous
1%
methylcellulose.
These
doses
were
selected
by
the
authors
to
correspond
to
30%
and
100%
of
the
calculated
MTD
for
this
compound.
Analysis
of
hepatocytes
for
unscheduled
DNA
synthesis
was
conducted
2
and
14
hours
after
treatment.
There
was
no
evidence
of
increased
DNA
synthesis
in
hepatocytes
from
rats
treated
with
any
tested
dose
of
bromoform.
Draft
­
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or
Quote
February
20,
2003
V
­
86
Table
V­
12
Summary
of
Mutagenicity,
Genotoxicity,
and
Neoplastic
Transformation
Data
for
Bromoform
Endpoint
Assay
System
Results
(
with/
without
activation)
d
References
In
Vitro
Studies
Gene
mutation
Salmonella
typhimurium
TA100a,
TA1535
NT/+
Simmon
and
Tardiff
(
1978)

TA1535,
TA1537b
TA100
TA97,
TA98
­/­
­/
±
c
±
c/­
NTP
(
1989a)

TA100b
­/+
Ishidate
et
al.
(
1982)

TA98
TA100,
TA1538a
+/+
­/­
Zeiger
(
1990)

S.
typhimurium
Ara
­/+
c
Roldan­
Arjona
and
Pueyo
(
1993)

RSJ100
NT/+
DeMarini
et
al.
(
1997)

Mouse
lymphoma
cells
b
+/+
NTP
(
1989a)

Chromosome
aberration
Chinese
hamster
fibroblasts
b
+/­
Ishidate
et
al.
(
1982)

Chinese
hamster
ovary
cells
b
­/
±
NTP
(
1989a)

DNA
damage
Human
lymphocytes
NT/+
Landi
et
al.
(
1999)

Sister
chromatid
exchange
Toadfish
leukocytes
a
NT/­
Maddock
and
Kelly
(
1980)

Human
lymphocytes
b
NT/+
Morimoto
and
Koizumi
(
1983)

Chinese
hamster
ovary
cells
b
­/
±
NTP
(
1989a)

Initiation
Rat
liver
GGT­
foci
assay
­
Herren­
Freund
and
Pereira
(
1986)

In
Vivo
Studies
Micronuclei
Mouse
bone
marrow
cells
­
Ishidate
et
al.
(
1982)

Mouse
bone
marrow
cells
­
Hayashi
et
al.
(
1988)

Mouse
bone
marrow
cells
­
Stocker
et
al.
(
1997)

Chromosome
aberrations
Mouse
bone
marrow
cells
­
NTP
(
1989a)

Rat
bone
marrow
cells
(
oral)
+
Fujie
et
al.
(
1990)
Table
V­
12
(
cont.)

Endpoint
Assay
System
Results
(
with/
without
activation)
d
References
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
87
DNA
damage
Rat
renal
cells
­
Potter
et
al.
(
1996)

Unscheduled
DNA
synthesis
Rat
hepatocytes
­
Stocker
et
al.
(
1997)

Sister
chromatid
exchange
Mouse
bone
marrow
cells
+
Morimoto
and
Koizumi
(
1983)

Mouse
bone
marrow
cells
(
ip)
+
NTP
(
1989a)

Sex­
linked
recessive
lethal
mutations
Drosophila
+
NTP
(
1989a)

NT
=
Not
Tested
a
Assay
was
conducted
in
a
closed
system.
b
Authors
did
not
specify
whether
or
not
the
assay
was
conducted
in
a
closed
system.
c
Equivocal
results
obtained.
d
With/
without
activation
applies
to
in
vitro
assays
only.

G.
Carcinogenicity
1.
Bromodichloromethane
NTP
(
1987)
evaluated
the
and
carcinogenic
potential
of
bromodichloromethane
in
F344/
N
rats
in
a
two­
year
study
oral
exposure
study.
Additional
details
of
this
study
are
provided
in
Section
V.
D.
1.
Groups
of
male
and
female
rats
(
50/
sex/
group)
were
administered
bromodichloromethane
in
corn
oil
via
gavage
at
doses
of
0,
50,
or
100
mg/
kg­
day
for
5
days/
week
for
102
weeks.
All
animals
were
examined
grossly
and
microscopically
for
neoplastic
lesions.
Survival
of
all
dosed
animals
was
comparable
to
or
greater
than
the
corresponding
control
group.
Mean
body
weights
of
high­
dose
make
and
female
rats
was
decreased
during
the
last
1.5
years
of
the
study.
Body
weight
gains
of
high
dose
male
and
female
rats
were
86%
and
70%
of
the
corresponding
vehicle
control
group,
respectively.
Statistically
significant
increases
in
the
incidences
of
neoplasms
of
the
large
intestine
and
kidney
were
observed
in
male
and
female
rats
(
Table
V­
13).
The
study
authors
noted
that
neoplasms
of
the
large
intestine
and
kidney
are
uncommon
tumors
in
F344/
N
rats
based
on
historical
control
data
for
NTP
studies.
They
concluded
that
under
the
conditions
of
these
2­
year
gavage
studies,
clear
evidence
of
carcinogenic
activity
existed
in
male
and
female
rats.

NTP
(
1987)
also
evaluated
the
potential
toxic
and
carcinogenic
effects
of
bromodichloromethane
in
rats
and
mice
in
a
two­
year
study
oral
exposure
study.
Additional
details
of
this
study
are
provided
in
Section
V.
D.
1.
Groups
of
male
and
female
B6C3F
1
mice
Table
V­
12
(
cont.)

Draft
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February
20,
2003
V
­
88
(
50/
sex/
dose)
were
administered
doses
of
0,
25,
or
50
mg/
kg­
day
(
males)
or
0,
75,
or
150
mg/
kgday
(
females)
for
5
days/
week
for
102
weeks.
All
animals
were
examined
grossly
and
Draft
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February
20,
2003
V
­
89
Table
V­
13
Tumor
Frequencies
in
F344/
N
Rats
and
B6C3F1
Mice
Exposed
to
Bromodichloromethane
in
Corn
Oil
for
2
Years
­
Adapted
from
NTP
(
1987)

Animal
Tissue/
Tumor
Tumor
Frequency
Control
50
mg/
kg
100
mg/
kg
Male
Rat
Large
intestine
a
Adenomatous
polyp
0/
50
3/
49
33/
50b
Adenocarcinoma
0/
50
11/
49b
38/
50b
Combined
0/
50
13/
49b
45/
50b
Kidney
a
Tubular
cell
adenoma
0/
50
1/
49
3/
50
Tubular
cell
adenocarcinoma
0/
50
0/
49
10/
50b
Combined
0/
50
1/
49
13/
50b
Control
50
mg/
kg
100
mg/
kg
Female
Rat
Large
intestine
c
Adenomatous
polyp
0/
46
0/
50
7/
47b
Adenocarcinoma
0/
46
0/
50
6/
47b
Combined
0/
46
0/
50
12/
47b
Kidney
Tubular
cell
adenoma
0/
50
1/
50
6/
50b
Tubular
cell
adenocarcinoma
0/
50
0/
50
9/
50b
Combined
0/
50
1/
50
15/
50b
Control
25
mg/
kg
50
mg/
kg
Male
Mouse
Kidney
d
Tubular
cell
adenoma
1/
46
2/
49
6/
50
Tubular
cell
adenocarcinoma
0/
46
0/
49
4/
50
Combined
1/
46
2/
49
9/
50b
Control
75
mg/
kg
150
mg/
kg
Female
Mouse
Liver
Hepatocellular
adenoma
1/
50
13/
48b
23/
50b
Hepatocellular
carcinoma
2/
50
5/
48
10/
50b
Combined
3/
50
18/
48b
29/
50b
a
One
rat
died
at
week
33
in
the
low­
dose
group
and
was
eliminated
from
the
cancer
risk
calculation.
b
Statistically
significant
at
p<
0.05,
compared
to
controls.
c
Intestine
not
examined
in
four
rats
from
control
group
and
three
rats
from
high­
dose
group.
d
In
the
control
group,
two
mice
died
during
the
first
week,
one
mouse
died
during
week,
nine
and
one
escaped
in
week
79.
One
mouse
in
the
low­
dose
group
died
in
the
first
week.
All
of
these
mice
were
eliminated
from
the
cancer
risk
calculations.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
90
microscopically
for
neoplastic
lesions.
Survival
of
dosed
male
mice
was
comparable
to
the
corresponding
control
group.
Survival
of
dosed
and
vehicle
control
females
was
decreased
after
week
84
as
a
result
of
ovarian
abscesses.
Body
weight
gain
in
high­
dose
males
was
decreased
by
13%
when
compared
to
the
vehicle
control
group.
Body
weight
gain
in
low­
and
high­
dose
females
was
reduced
by
25%
and
55%,
respectively.
Statistically
significant
increases
were
observed
in
the
incidences
of
neoplasms
of
the
kidney
in
male
mice
and
the
liver
in
female
mice
(
Table
V­
13).
The
study
authors
noted
that
neoplasms
of
the
kidney
are
uncommon
in
B6C3F
1
mice
based
on
NTP
historical
control
data.
They
concluded
that
under
the
conditions
of
these
2­
year
gavage
studies,
clear
evidence
of
carcinogenic
activity
existed
in
male
and
female
mice.

Tumasonis
et
al.
(
1987)
exposed
groups
of
58
male
and
female
Wistar
rats
to
bromodichloromethane
in
drinking
water
from
weaning
until
death
occurred
in
all
of
the
animals
(
approximately
185
weeks).
The
exposure
level
was
2,400
mg/
L
for
72
weeks
and
was
reduced
to
1,200
mg/
L
for
the
remaining
113
weeks.
Based
on
a
graph
presented
by
the
authors,
the
average
dose
over
the
course
of
the
experiment
was
probably
about
150
mg/
kg­
day
for
females
and
about
100
mg/
kg­
day
for
males.
Exposed
animals
of
both
sexes
gained
significantly
less
weight
(
approximately
30
to
40%)
than
control
animals.
There
was
a
statistically
significant
(
p
<
0.01)
increase
in
the
incidence
of
hepatic
neoplastic
nodules
in
exposed
females
compared
to
control
females
(
32%
versus
0%).
Significant
increases
were
also
reported
for
the
occurrence
of
hepatic
adenofibrosis
(
12%
versus
0%)
and
lymphosarcoma
(
17%
versus
11%)
in
females.
No
statistically
significant
increase
in
the
incidence
of
any
tumor
was
reported
in
males.
Two
males
and
one
female
among
the
treated
animals
were
observed
to
have
renal
adenoma
or
carcinoma,
while
no
renal
tumors
were
observed
in
the
controls.
Statistically
significant
decreases
in
the
incidence
of
mammary
tumors
and
pituitary
tumors
in
females
and
lymphosarcomas
in
males
were
observed.

Aida
et
al.
(
1992b)
administered
bromodichloromethane
to
Slc:
Wistar
rats
(
40/
sex/
treatment
group
and
70/
sex/
controls)
at
dietary
levels
of
0%,
0.014%,
0.055%,
or
0.22%
for
up
to
24
months.
The
test
material
was
microencapsulated
and
mixed
with
powdered
feed.
Based
on
the
mean
food
intakes,
the
mean
doses
were
0,
6.1,
25.5,
or
138.0
mg/
kg­
day
for
males
and
0,
8.0,
31.7,
or
168.4
mg/
kg­
day
for
females.
The
only
neoplastic
lesions
observed
were
three
cholangiocarcinomas
and
two
hepatocellular
adenomas
in
the
high­
dose
females,
one
hepatocellular
adenoma
in
a
control
female,
one
cholangiocarcinoma
in
a
high­
dose
male,
and
one
hepatocellular
adenoma
each
in
a
low­
dose
male
and
a
high­
dose
male.
Based
on
these
results,
the
study
authors
concluded
that
there
was
no
clear
evidence
that
microencapsulated
bromodichloromethane
administered
in
the
diet
was
carcinogenic
in
Wistar
rats.

Voronin
et
al.
(
1987)
assessed
the
carcinogenic
potential
of
bromodichloromethane
in
male
and
female
CBA
x
C57Bl/
6
mice.
Groups
of
mice
(
50­
55/
sex/
concentration)
were
exposed
to
bromodichloromethane
provided
in
drinking
water
at
concentrations
of
0.04,
4.0,
or
400
mg/
L
for
104
weeks.
Untreated
control
groups
of
75
male
and
50
female
mice
were
also
included
in
the
study
design.
No
significant
differences
were
observed
in
total
tumor
incidence
when
evaluated
by
Chi
square
analysis.
The
study
authors
concluded
that,
under
the
conditions
of
this
bioassay,
bromodichloromethane
was
not
carcinogenic
in
mice.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
91
Theiss
et
al.
(
1977)
examined
the
carcinogenic
potential
of
bromodichloromethane
in
Strain
A
mice
(
6
to
8
weeks
old).
Male
animals
(
20
mice/
group)
were
injected
intraperitoneally
up
to
three
times
weekly
over
a
period
of
8
weeks.
Three
dose
levels
(
20,
40,
or
100
mg/
kg
bromodichloromethane)
were
used
with
concurrent
positive
and
negative
control
groups
that
contained
20
animals
each.
Mice
were
sacrificed
24
weeks
after
the
first
injection,
and
the
frequency
of
lung
tumors
in
each
test
group
was
compared
with
vehicle­
treated
controls.
No
statistically
significant
increase
in
the
incidence
of
lung
tumors/
mouse
was
reported.

Melnick
et
al.
(
1998)
investigated
the
mechanistic
relationship
between
liver
toxicity
and
tumorigenicity
of
bromodichloromethane.
Female
B6C3F
1
mice
(
10
animals
per
group)
were
treated
with
bromodichloromethane
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
Doses
of
bromodichloromethane
used
in
this
study
were
0
(
vehicle
only),
75,
150,
or
326
mg/
kg­
day.
A
significant
dose­
related
increase
in
absolute
liver
weight
and
liver
weight/
body
weight
ratio
was
noted
for
the
150
and
326
mg/
kg­
day
dose
groups.
Serum
ALT
activity
was
significantly
increased
in
the
two
highest
dose
groups
and
serum
SDH
activity
was
elevated
at
all
doses
tested.
At
necropsy,
there
was
clear
evidence
of
hepatocyte
hydropic
degeneration
in
animals
treated
with
150
and
326
mg/
kg­
day.
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
The
two
highest
(
150
and
326
mg/
kg­
day)
doses
resulted
in
significantly
elevated
hepatocyte
proliferation
as
measured
by
the
LI.
Using
the
Hill
equation
model,
these
authors
compared
the
dose
response
for
liver
toxicity
(
enzyme
and
LI
data)
and
tumorigenicity
(
data
from
previously
published
NTP
sponsored
bioassays)
for
bromodichloromethane.
This
analysis
indicated
that
the
shape
of
the
dose
response
as
well
as
the
Hill
exponents
were
different
for
liver
toxicity
and
tumorigenicity.
It
was
concluded
that
the
results
of
this
comparison
do
not
support
a
causal
relationship
between
liver
toxicity/
reparative
hyperplasia
and
tumor
development.

George
et
al.
(
2002)
evaluated
the
carcinogenicity
of
bromodichloromethane
in
male
F344/
N
rats
(
78
animals/
dose)
exposed
to
the
compound
via
drinking
water
for
104
weeks.
Nominal
concentrations
of
0.07,
0.35,
or
0.70
g/
L
were
administered
in
drinking
water
containing
0.25%
Emulphor
®
.
The
vehicle
control
solution
consisted
of
0.25%
Emulphor
®
.
The
study
authors
indicated
that
testing
of
higher
concentrations
was
prevented
by
refusal
of
the
test
animals
to
drink
solutions
containing
more
than
0.7
g/
L.
Six
animals
per
exposure
concentration
were
sacrificed
at
13,
26,
52,
and
78
weeks
for
gross
observation
and
histopathological
examination
of
the
thyroid,
liver,
stomach,
duodenum,
jejunum,
ileum,
colon,
rectum,
spleen,
kidneys,
urinary
bladder,
and
testes.
A
complete
rodent
necropsy
was
performed
at
terminal
sacrifice
and
representative
samples
of
the
tissues
listed
above
were
examined
microscopically.
A
complete
pathological
examination
was
performed
on
five
rats
from
the
high
dose
group.
Serum
profiles
of
LDH,
ALT,
ALP,
AST,
SDH,
BUN,
total
protein,
creatine,
and
total
antioxidant
activities
were
determined
at
26,
52,
and
104
weeks.
Hepatocyte
and
renal
tubular
cell
proliferation
were
measured
at
each
sacrifice
by
bromodeoxyuridine
labeling.

The
measured
drinking
water
concentrations
of
bromodichloromethane
were
0.06,
0.38,
and
0.76
g/
L.
When
corrected
for
loss
of
bromodichloromethane
as
a
result
of
volatility,
instability,
or
adsorption
to
glass
surfaces
during
treatment,
the
corresponding
administered
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
92
concentrations
were
0.06,
0.33,
and
0.62
g/
L.
Based
on
measured
water
consumption,
these
levels
correspond
to
mean
daily
doses
for
the
entire
study
of
3.9,
20.6,
and
36.3
mg/
kg­
day
as
calculated
by
the
study
authors.
Mean
daily
doses
of
6.4,
32.6,
and
58.9
mg/
kg­
day
were
calculated
for
the
first
13
weeks
of
the
study
when
the
growth
rate
of
the
test
animals
was
highest.
No
significant
differences
were
observed
among
groups
for
feed
consumption
or
survival.
Twenty­
one
to
22
unscheduled
deaths
were
observed
in
each
treatment
group.
Mononuclear
cell
leukemia
was
seen
in
all
dose
groups
and
was
reported
to
be
the
primary
cause
of
morbidity
and
mortality
prior
to
104
weeks.
Exposure
to
bromodichloromethane
did
not
affect
the
growth
rate
of
test
animals
when
compared
to
the
control.
Kidney
weight
was
significantly
depressed
at
the
high
dose
and
a
significant
negative
trend
was
observed
for
relative
kidney
weight.
No
significant
changes
were
observed
in
clinical
chemistry
parameters.
Observed
nonneoplastic
changes
in
the
liver
(
e.
g.,
biliary
fibrosis,
bile
duct
inflammation,
and
chronic
inflammation)
were
considered
to
be
age­
related
background
changes,
since
neither
the
incidence
nor
severity
of
the
lesions
differed
from
the
control
values.
Bromodichloromethane
had
no
effect
on
hepatocyte
proliferation
as
measured
by
bromodeoxyuridine
labeling.
Renal
tubular
cell
hyperplasia
was
significantly
decreased
in
the
3.9
mg/
kg­
day
group
and
significantly
increased
in
the
36.3
mg/
kg­
day
group
(
15.8%)
relative
to
the
control
value
(
8.7%).

The
absence
of
effect
on
body
weight
and
other
examined
endpoints
suggests
that
a
maximum
toxic
dose
may
not
have
been
achieved
in
this
study.
However,
the
dosing
regimen
used
by
George
et
al.
(
2002)
was
sufficient
to
increase
the
incidence
of
hepatocellular
neoplasia
(
Table
V­
14).
The
data
for
hepatic
tumors
indicate
a
biphasic
pattern
of
dose­
response.
The
prevalence
and
multiplicity
of
hepatocellular
adenoma
and
combined
hepatocellular
adenoma
and
carcinoma
were
significantly
increased
at
3.9
mg/
kg­
day,
nonsignificantly
increased
at
20.6
mg/
kg­
day,
and
comparable
to
the
control
values
at
36.3
mg/
kg­
day.
The
prevalence
and
multiplicity
of
hepatocellular
carcinoma
were
increased
at
20.6
mg/
kg­
day
when
compared
to
control
values,
but
the
response
did
not
reach
statistical
significance.
The
underlying
basis
for
the
biphasic
response
is
unknown,
but
the
study
authors
noted
that
the
observed
pattern
of
response
could
be
explained
by
inhibition
of
the
hepatic
metabolism
of
bromodichloromethane
by
the
compound
itself.
Exposure
to
bromodichloromethane
decreased
the
prevalence
of
basophilic
(
control,
67%;
3.6
mg/
kg­
day,
62.2%;
20.6
mg/
kg­
day,
46%;
36.6
mg/
kg­
day,
34.7%)
and
clear
cell
(
17.8%,
2.2%,
2.1%,
4.1%)
altered
foci
of
cells
(
AFC)
in
a
dose­
dependent
manner,
but
had
no
significant
effect
on
the
prevalence
of
eosinophilic
AFCs
when
compared
to
the
controls.
The
decreases
in
prevalence
were
statistically
significant
at
the
mid
and
high
doses
for
basophilic
AFCs
and
at
all
doses
for
clear
cell
AFCs.
Exposure
to
bromodichloromethane
had
no
significant
effect
on
the
prevalence
of
renal
tubular
adenomas
or
carcinomas
(
Table
V­
14).
One
renal
tubular
adenoma
was
observed
in
the
3.6
mg/
kg­
day
group
and
two
tumors
were
observe
in
the
36.3
mg/
kg­
day.
The
historical
incidence
of
renal
tubular
adenomas
in
male
F344/
N
rats
is
very
low
(
2/
327
or
0.6%),
as
determined
from
control
groups
in
NTP
drinking
water
studies
(
NTP,
2002;
no
data
were
reported
from
the
study
laboratory).
Therefore,
the
occurrence
of
these
tumors
in
the
present
study
may
be
of
biological
significance.
No
increased
incidences
of
neoplasia
were
evident
in
the
five
high
dose
animals
selected
for
a
histopathological
examination
of
all
organs.
On
the
basis
of
the
increased
prevalence
and
multiplicity
of
hepatocellular
neoplasms
in
the
3.9
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
93
and
20.6
mg/
kg­
day
groups,
the
study
authors
concluded
that
bromodichloromethane
was
carcinogenic
in
male
F344/
N
rats
under
the
Table
V­
14
Hepatic
and
Renal
Tumors
in
Male
F344/
N
Rats
Administered
Bromodichloromethane
in
the
Drinking
Water
for
Two
Years
Tumor
Type
Mean
Daily
Dose
of
Bromodichloromethane
(
mg/
kg­
day)

Vehicle
Control
3.9
20.6
36.3
Liver
Hepatocellular
adenoma
1/
45
(
2.2%)
a
0.02
±
0.02b,
c
7/
45
(
15.5%)*
0.16
±
0.04*
3/
48
(
6.2%)
0.06
±
0.02
2/
49
(
4.1%)
0.04
±
0.02
Hepatocellular
carcinoma
1/
45
(
2.2%)
0.02
±
0.02
1/
45
(
2.2%)
0.02
±
0.01
4/
48
(
8.3%)
0.10
±
0.03
2/
49
(
4.1%)
0.04
±
0.02
Hepatocellular
adenoma
and
carcinoma
(
combined)
2/
45
(
4.4%)
0.04
±
0.02
8/
45
(
17.8%)*
0.19
±
0.00*
7/
48
(
14.6%)
0.17
±
0.04
4/
49
(
8.2%)
0.08
±
0.28
Kidney
Tubular
cell
adenoma
0/
46
(
0%)
1/
45
(
2.2%)
0/
51
(
0%)
2/
44
(
4.5%)

Tubular
cell
carcinoma
0/
46
(
0%)
0/
45
(
0%)
0/
51
(
0%)
0/
44
(
0%)

Tubular
cell
adenoma
or
carcinoma
(
combined)
0/
46
(
0%)
1/
45
(
2.2%)
0/
51
(
0%)
2/
44
(
4.5%)

Source:
George
et
al.
(
2002)
*
Statistically
significant
when
compared
to
the
control
value,
p

0.05
a
Prevalence
(
percentage
of
animals
with
tumor)
b
Multiplicity,
number
of
tumors
per
animal
c
Mean
±
standard
deviation
conditions
of
the
bioassay.
A
source
of
uncertainty
in
this
conclusion
is
lack
of
knowledge
on
the
biological
mechanism
underlying
the
biphasic
dose­
response
observed
for
hepatic
tumors.

George
et
al.
(
2002)
also
evaluated
the
carcinogenicity
of
bromodichloromethane
in
male
B6C3F
1
mice
(
78
animals/
dose)
exposed
via
drinking
water
for
100
weeks.
Nominal
concentrations
of
0.05,
0.25,
or
0.50
g/
L
were
administered
in
drinking
water
containing
0.25%
Emulphor
®
.
The
vehicle
control
solution
consisted
of
0.25%
Emulphor
®
.
Seven
animals
per
exposure
concentration
were
sacrificed
at
13,
26,
52,
and
78
weeks
for
gross
observation
and
histopathological
examination
of
the
liver,
stomach,
duodenum,
jejunum,
ileum,
colon,
rectum,
spleen,
kidneys,
urinary
bladder,
and
testes.
A
complete
rodent
necropsy
was
performed
at
Draft
­
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or
Quote
February
20,
2003
V
­
94
terminal
sacrifice
and
representative
samples
of
the
tissues
listed
above
were
examined
microscopically.
A
complete
pathological
examination
was
performed
on
five
rats
from
the
high
dose
group.
Serum
profiles
of
LDH,
ALT,
ALP,
AST,
SDH,
BUN,
total
protein,
creatine,
and
total
antioxidant
activities
were
determined
at
26,
52,
and
100
weeks.
Hepatocyte
and
renal
tubular
cell
proliferation
were
measured
by
bromodeoxyuridine
labeling
at
each
sacrifice.

The
measured
drinking
water
concentrations
of
bromodichloromethane
were
0.06,
0.30,
and
0.55
g/
L.
When
corrected
for
loss
of
bromodichloromethane
as
a
result
of
volatility,
instability,
or
adsorption
to
glass
surfaces
during
treatment,
the
corresponding
administered
concentrations
were
0.06,
0.28,
and
0.49
g/
L.
Based
on
measured
water
consumption,
these
levels
correspond
to
mean
daily
doses
of
8.1,
27.2,
and
43.4
mg/
kg­
day
as
calculated
by
the
study
authors.
Water
consumption
was
significantly
reduced
at
the
mid­
and
high
doses;
the
study
authors
attributed
the
reduced
intake
to
taste
aversion.
No
significant
differences
were
observed
among
groups
for
feed
consumption
or
survival.
Exposure
to
bromodichloromethane
did
not
affect
the
growth
rate
of
test
animals
when
compared
to
the
control.
Kidney
weight
was
significantly
depressed
at
27.2
and
43.4
mg/
kg­
day
when
compared
to
the
control
values.
No
significant
changes
were
observed
in
clinical
chemistry
parameters.
Mild,
treatment­
related
nonneoplastic
hepatic
lesions
were
observed
in
the
27.2
and
43.4
mg/
kg­
day
dose
groups
(
identity
and
prevalence
not
reported).
Increased
incidences
of
hepatocellular
karyomegaly
and
necrosis
with
inflammation
(
prevalence
and
severity
not
reported)
were
not
dose­
related.
The
prevalence
of
renal
tubular
hyperplasia
was
3%,
0%,
6%
and
0%
for
the
vehicle
control,
8.1,
27.2,
and
43.4
mg/
kg­
day
groups,
respectively.
Other
observed
preneoplastic
and
neoplastic
lesions
(
identity
and
prevalence
not
reported)
were
considered
background
events
for
the
male
B6C3F
1
mouse.
BrdU
labeling
index
in
hepatocytes
and
renal
tubular
cells
was
not
altered
at
any
time
point.
Hepatocellular
adenomas
and
carcinomas
were
observed
in
all
treatment
groups.
Neither
the
prevalence
nor
multiplicity
of
these
tumors
was
significantly
increased
by
exposure
to
bromodichloromethane.
Renal
tubular
cell
neoplasia
was
not
observed
in
any
treatment
group.
No
increased
incidences
of
neoplasia
were
evident
in
the
five
high
dose
animals
subject
to
a
full
histopathological
examination.
On
the
basis
of
these
data,
the
study
authors
concluded
that
bromodichloromethane
was
not
carcinogenic
to
male
mice
under
the
conditions
employed
in
this
study.
However,
in
the
absence
of
compound­
related
effects
on
body
weight
or
other
toxicologic
endpoints,
it
is
not
evident
that
an
adequately
high
dose
was
tested
in
this
study.

De
Angelo
et
al.,
(
2002)
evaluated
the
ability
of
bromodichloromethane
administered
in
drinking
water
to
induce
aberrant
crypt
foci
(
ACF),
putative
early
preneoplastic
lesions,
in
the
colons
of
male
F344/
N
rats.
Groups
of
weanling
rats
(
6
animals/
group)
were
exposed
to
distilled
water,
0.25%
Alkamuls
EL­
620
®
,
or
0.7
g/
L
bromodichloromethane
in
0.25%
Alkamuls
EL­
620
for
13
weeks.
A
single
intraperitoneal
injection
of
30
mg/
kg
azoxymethane
(
AOM)
served
as
the
positive
control.
Body
weight
and
water
consumption
were
measured
twice
during
the
first
week
of
the
study
and
once
per
week
thereafter.
Colons
were
collected
at
study
termination,
fixed,
stained
with
0.2%
methylene
blue,
divided
into
three
equal
segments,
and
scanned
for
ACF.
The
measured
concentration
of
bromodichloromethane
averaged
0.64
±
0.06
mg/
L
(
mean
and
standard
error)
over
the
course
of
the
study.
When
adjusted
for
volatilization
and
adherence
to
glass,
the
corrected
concentration
was
0.51
mg/
L.
Water
consumption
was
significantly
reduced
Draft
­
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or
Quote
February
20,
2003
V
­
95
(
38%)
in
the
bromodichloromethane
exposure
group
when
compared
to
the
deionized
water
control.
The
average
daily
dose
was
45
mg/
kg­
day
as
calculated
by
the
study
authors.
Average
terminal
body
weight
of
the
rats
exposed
to
bromodichloromethane
was
within
10%
of
the
control
values.
No
ACF
were
observed
in
colons
from
control
animals.
ACF
were
observed
in
five
of
six
colons
from
bromodichloromethane
exposed
animals.
The
total
number
of
ACF
(
30)
and
number
of
aberrant
crypts
per
focus
(
3.33
±
0.47)
were
significantly
increased
relative
to
the
combined
deionized
water
and
vehicle
controls.
All
observed
ACF
were
located
in
the
distal
(
rectal)
segment
of
the
colon.
For
comparison,
807
ACF
and
4.95
±
0.25
crypts
per
focus
(
mean
and
standard
error)
were
observed
in
the
AOM
positive
control
group.
Eight
percent,
42%
and
50%
of
the
ACF
induced
by
AOM
were
located
in
the
proximal,
middle,
and
distal
segment
of
the
colon,
respectively.
The
study
authors
noted
that
ACF
induced
by
bromodichloromethane
do
not
to
progress
to
neoplasia,
as
judged
by
the
absence
of
colon
neoplasms
in
the
two­
year
cancer
study
conducted
by
George
et
al.
(
2002).

De
Angelo
et
al.
(
2002)
also
evaluated
the
ability
of
bromodichloromethane
administered
in
drinking
water
to
induce
ACF
in
the
colons
of
male
B6C3F
1
(
6
animals/
group)
and
A/
J
mice
(
9
animals/
group;
sex
not
specified).
Mice
of
the
A/
J
strain
are
sensitive
to
chemical
induction
of
ACF.
Test
animals
were
exposed
to
distilled
water,
0.25%
Alkamuls
EL­
620
®
,
or
0.5
g/
L
bromodichloromethane
in
0.25%
Alkamuls
EL­
620
for
13
weeks
(
both
strains)
or
30
weeks
(
A/
J
mice
only).
A
single
intraperitoneal
injection
of
50
mg/
kg
4­
aminobiphenyl
or
10
mg/
kg
azoxymethane
(
AOM)
served
as
the
positive
controls
for
the
B6C3F1
and
A/
J
strains,
respectively.
Body
weight
and
water
consumption
were
measured
twice
during
the
first
week
of
the
study
and
once
per
week
thereafter.
Colons
were
collected
at
study
termination,
fixed,
stained
with
0.2%
methylene
blue,
divided
into
three
equal
segments,
and
scanned
for
ACF.
The
study
report
did
not
provide
results
for
measured
concentration
of
bromodichloromethane
in
drinking
water
solutions
or
an
estimated
dose.
No
differences
were
observed
in
between
the
control
and
any
treatment
group
for
body
weight
or
water
and
feed
consumption.
ACF
development
was
not
observed
in
the
colons
of
B6C3F
1
mice
treated
with
bromodichloromethane
in
the
drinking
water
or
injected
with
4­
aminobiphenyl.
Bromodichloromethane
did
not
induce
ACF
in
A/
J
mice.
Injection
of
A/
J
mice
with
the
positive
control
compound
AOM
induced
47.4
±
4.9
ACF/
cm2
(
mean
and
standard
error)
and
7.2
±
1.1
tumors/
cm2
after
13
weeks
and
17.8
±
2.6
tumors/
cm2
after
30
weeks
of
treatment.
In
comparison,
807
ACF
and
4.95
±
0.25
crypts
per
focus
were
observed
in
the
AOM
positive
control
group.

2.
Dibromochloromethane
NTP
(
1985)
administered
dibromochloromethane
at
doses
of
0,
40,
or
80
mg/
kg­
day
(
in
corn
oil)
to
groups
of
50
male
and
50
female
F344/
N
rats
via
gavage
5
times/
week
for
104
to
105
weeks.
Survival
of
dosed
male
and
female
rats
was
comparable
to
that
of
the
vehicle­
control
groups.
High­
dose
males
had
lower
body
weights
when
compared
with
the
vehicle
control.
Compound­
related
nonneoplastic
lesions
(
fatty
metamorphosis
and
ground­
glass
cytoplasmic
changes)
were
found
in
the
livers
of
both
sexes
(
See
section
V.
D.
1).
Nephrosis
was
observed
in
female
rats.
No
statistically
significant
increase
in
the
incidence
of
any
neoplastic
lesion
was
Draft
­
Do
Not
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or
Quote
February
20,
2003
V
­
96
observed.
Based
on
the
results
of
this
study,
the
authors
concluded
that
there
was
no
evidence
of
carcinogenicity
in
rats
administered
dibromochloromethane.

NTP
(
1985)
administered
dibromochloromethane
(
in
corn
oil)
to
groups
of
50
male
and
50
female
B6C3F
1
mice
via
gavage
5
times/
week
for
104
to
105
weeks.
The
administered
doses
were
0,
50,
or
100
mg/
kg­
day.
Survival
of
female
mice
was
comparable
to
that
of
the
vehiclecontrol
group.
High­
dose
male
mice,
however,
had
lower
survival
rates
than
the
vehicle
controls.
At
week
82,
nine
high­
dose
male
mice
died
of
an
unknown
cause.
An
inadvertent
overdose
of
dibromochloromethane
given
to
low­
dose
male
and
female
mice
at
week
58
killed
35
male
mice,
but
apparently
did
not
affect
the
females.
The
low­
dose
male
mouse
group
was,
therefore,
considered
to
be
unsuitable
for
analysis
of
neoplasms.
Compound­
related
nonneoplastic
lesions
were
found
primarily
in
the
livers
of
male
mice
(
hepatocytomegaly,
necrosis,
fatty
metamorphosis)
and
female
mice
(
calcification
and
fatty
metamorphosis).
Nephrosis
was
observed
in
male
mice.
In
females,
a
statistically
significant
increase
in
the
incidence
of
hepatocellular
adenomas
and
adenomas
and
carcinomas
combined
was
observed
in
the
high­
dose
group.
In
male
mice,
a
statistically
significant
increase
in
the
incidence
of
hepatocellular
carcinomas
and
adenomas
and
carcinomas
combined
was
observed
in
the
high­
dose
group.
A
summary
of
the
incidence
of
these
tumors
is
presented
in
Table
V­
15.
A
negative
trend
in
the
incidence
of
malignant
lymphomas
was
evident
in
dibromochloromethane­
exposed
male
mice
when
compared
to
the
vehicle
control.
The
study
authors
concluded
that
the
results
of
this
study
provided
equivocal
evidence
of
dibromochloromethane
carcinogenicity
in
male
B6C3F
1
mice
and
some
evidence
of
carcinogenicity
in
female
B6C3F
1
mice.

In
other
bioassays,
Voronin
et
al.
(
1987)
observed
no
significant
tumor
increases
in
CBAxC57B1/
6
mice
(
50/
sex/
dose)
treated
with
dibromochloromethane
in
the
drinking
water
at
concentrations
of
0,
0.04,
4.0,
or
400
mg/
L
(
approximately
0,
0.008,
0.76,
or
76
mg/
kg­
day)
for
104
weeks.
In
an
unpublished
report
of
a
two­
year
dietary
study,
Tobe
et
al.
(
1982)
reported
no
increase
in
gross
tumors
in
male
rats
dosed
with
up
to
210
mg/
kg­
day
or
female
rats
treated
with
up
to
350
mg/
kg­
day.

Melnick
et
al.
(
1998)
exposed
female
B6C3F
1
mice
(
10/
dose)
to
dibromochloromethane
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
The
doses
of
dibromochloromethane
in
this
study
were
0
(
vehicle
only),
50,
100,
192,
or
417
mg/
kg­
day.
The
corresponding
time­
weighted
doses
are
0,
37,
71,
137,
and
298
mg/
kg­
day.
No
treatment­
related
signs
of
overt
toxicity
were
observed
during
the
study.
Body
weight
and
water
intake
were
not
significantly
altered
at
any
dose
tested.
However,
a
statistically
significant
and
dose­
related
increase
in
liver
weight/
body
weight
ratio
was
seen
in
the
100,
192
and
417
mg/
kg­
day
dose
groups.
Serum
ALT
activity
was
significantly
increased
in
the
two
highest
dose
groups.
The
activity
of
serum
SDH
was
significantly
elevated
at
all
doses
tested
except
50
mg/
kg­
day.
However,
the
increase
in
activity
(
shown
graphically)
was
very
small
relative
to
the
control
at
the
100
and
192
mg/
kg­
day
doses.
At
necropsy,
there
was
clear
evidence
of
hepatocyte
hydropic
degeneration
in
the192
and
417
mg/
kg­
day
dose
groups.
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
Only
the
highest
dose
tested
(
417
mg/
kg­
day)
resulted
in
significantly
elevated
hepatocyte
proliferation
as
measured
by
the
LI.
Draft
­
Do
Not
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or
Quote
February
20,
2003
V
­
97
Using
the
Hill
equation
model,
these
authors
compared
the
dose
response
for
liver
toxicity
(
enzyme
and
LI
data)
and
tumorigenicity
(
data
from
previously
conducted
NTP
bioassays)
for
dibromochloromethane.
This
analysis
indicated
that
the
shape
of
the
dose
response
as
well
as
the
Hill
exponents
were
different
for
liver
toxicity
and
tumorigenicity.
It
was
concluded
that
the
results
of
this
comparison
do
not
support
a
causal
relationship
between
liver
toxicity/
reparative
hyperplasia
and
tumor
development.

Table
V­
15
Frequencies
of
Liver
Tumors
in
B6C3F1
Mice
Administered
Dibromochloromethane
in
Corn
Oil
for
105
Weeks
­
Adapted
from
NTP
(
1985)

Treatment
(
mg/
kg­
day)
Sex
Adenoma
Carcinoma
Adenoma
or
Carcinoma
(
combined)

Vehicle
Control
M
F
14/
50
2/
50
10/
50
4/
50
23/
50
6/
50
50
M
F
­­
a
4/
49
­­
6/
49
­­
10/
49
100
M
F
10/
50
11/
50b
19/
50b
8/
50
27/
50c
19/
50d
a
Male
low­
dose
group
was
inadequate
for
statistical
analysis.
b
p
<
0.05
relative
to
controls.
c
p
<
0.01
(
life
table
analysis);
p
=
0.065
(
incidental
tumor
test)
relative
to
controls.
d
p
<
0.01
relative
to
controls.

De
Angelo
et
al.
(
2002)
evaluated
the
ability
of
dibromochloromethane
administered
in
drinking
water
to
induce
aberrant
crypt
foci
(
ACF),
putative
early
preneoplastic
lesions,
in
the
colons
of
male
F344/
N
rats.
Groups
of
weanling
rats
(
6
animals/
group)
were
exposed
to
distilled
water,
0.25%
Alkamuls
EL­
620
®
,
or
0.9
g/
L
dibromochloromethane
in
0.25%
Alkamuls
EL­
620
for
13
weeks.
A
single
intraperitoneal
injection
of
30
mg/
kg
azoxymethane
(
AOM)
served
as
the
positive
control.
Body
weight
and
water
consumption
were
measured
twice
during
the
first
week
of
the
study
and
once
per
week
thereafter.
Colons
were
collected
at
study
termination,
fixed,
stained
with
0.2%
methylene
blue,
divided
into
three
equal
segments,
and
examined
for
ACF.
The
measured
concentration
of
dibromochloromethane
averaged
0.80
±
0.05
mg/
L
(
mean
and
standard
error)
over
the
course
of
the
study.
When
adjusted
for
volatilization
and
adherence
to
glass,
the
corrected
concentration
was
0.63
mg/
L.
Water
consumption
was
significantly
reduced
(
32%)
in
the
dibromochloromethane
exposure
group
when
compared
to
the
deionized
water
control.
The
average
daily
dose
of
dibromochloromethane
was
60
mg/
kg­
day
as
calculated
by
the
study
authors.
Average
terminal
body
weight
of
the
rats
exposed
to
dibromochloromethane
was
within
10%
of
the
control
values.
No
ACF
were
observed
in
colons
from
control
animals.
ACF
were
observed
in
three
of
six
colons
from
dibromochloromethane­
exposed
animals.
The
total
number
of
ACF
(
17)
and
number
of
aberrant
crypts
per
focus
(
2.43
±
0.61)
were
significantly
increased
relative
to
the
combined
deionized
water
and
vehicle
controls.
Fourteen
percent
of
the
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
98
observed
ACF
were
located
in
the
middle
segment
of
the
colon
and
86%
were
located
in
the
distal
(
rectal)
segment.
In
comparison,
807
ACF
and
4.95
±
0.25
crypts
per
focus
were
observed
in
the
AOM
positive
control
group.
Eight
percent,
42%
and
50%
of
the
ACF
induced
by
AOM
were
located
in
the
proximal,
middle,
and
distal
segment
of
the
colon,
respectively.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
99
3.
Bromoform
Groups
of
50
male
B6C3F
1
mice
were
exposed
to
bromoform
via
gavage
(
corn
oil)
at
doses
of
0,
50,
or
100
mg/
kg­
day
of
bromoform
for
103
weeks
(
5
days/
week)
(
NTP,
1989a).
Groups
of
50
female
mice
received
doses
of
0,
100,
or
200
mg/
kg­
day
bromoform
by
the
same
protocol.
At
termination,
all
animals
underwent
gross
necropsy
and
thorough
histological
examinations
of
tissues.
Survival
in
both
treated
female
groups
was
reduced;
however,
the
authors
attributed
this
reduction
in
survival
partly
to
utero­
ovarian
infection.
A
statistically
significant
increase
in
the
incidence
of
thyroid
follicular
cell
hyperplasia
was
noted
in
high­
dose
females;
however,
there
were
no
statistically
significant
increases
in
the
incidence
of
any
neoplastic
lesion
in
any
dose
group
compared
to
controls.
Based
on
the
results
of
this
study,
the
NTP
(
1989a)
concluded
there
was
no
evidence
of
carcinogenic
activity
of
bromoform
in
male
or
female
mice.

In
a
similar
experiment,
groups
of
50
male
and
50
female
F344/
N
rats
were
administered
bromoform
via
gavage
in
oil
at
doses
of
0,
100,
or
200
mg/
kg­
day
for
5
days/
week
for
103
weeks
(
NTP
1989a).
At
termination,
all
animals
were
necropsied,
and
a
thorough
histological
examination
of
tissues
was
performed.
Adenomatous
polyps
or
adenocarcinomas
of
the
large
intestine
were
noted
in
three
high­
dose
male
rats,
eight
high­
dose
female
rats,
and
one
low­
dose
female
rat
(
Table
V­
16).
Although
the
number
of
tumors
found
was
small,
the
incidence
was
considered
to
be
significant
because
these
intestinal
tumors
are
very
rare
in
the
rat.
The
NTP
concluded
that
there
was
some
evidence
for
carcinogenic
activity
in
male
rats
and
clear
evidence
in
female
rats.

Table
V­
16
Tumor
Frequencies
in
the
Large
Intestine
of
F344/
N
Rats
Exposed
to
Bromoform
in
Corn
Oil
for
2
Years
­
Adapted
from
NTP
(
1989a)

Tumor
Tumor
Frequency
Male
rat
Control
100
mg/
kg
200
mg/
kg
Adenocarcinoma
0/
50
0/
50
1/
50
Polyp
(
adenomatous)
0/
50
0/
50
2/
50
Female
rat
Control
100
mg/
kg
200
mg/
kg
Adenocarcinoma
0/
48
0/
50
2/
50
Polyp
(
adenomatous)
0/
48
1/
50
6/
50
Theiss
et
al.
(
1977)
examined
the
carcinogenic
activity
of
bromoform
in
Strain
A
mice.
Male
animals
(
6
to
8
weeks
old,
20
mice/
group)
were
injected
intraperitoneally
up
to
three
times
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
100
weekly
over
a
period
of
8
weeks.
The
dose
levels
utilized
were
4,
48,
or
100
mg/
kg
bromoform.
A
positive
and
a
negative
control
group
were
included
in
the
study
design
and
each
contained
20
animals.
Mice
were
sacrificed
24
weeks
after
the
first
injection
and
the
frequency
of
lung
tumors
in
each
test
group
was
compared
with
vehicle­
treated
controls.
Bromoform
produced
a
significant
increase
(
p
=
0.041)
in
tumor
frequency
only
at
the
intermediate
dose.
U.
S.
EPA
(
1980)
concluded
that
these
results
were
suggestive
of
carcinogenic
activity
but
were
not
an
adequate
basis
for
the
assessment
of
cancer
risk.

Kurokawa
(
1987)
observed
no
evidence
of
carcinogenicity
in
male
or
female
rats
exposed
to
microencapsulated
bromoform
at
concentrations
of
400,
1600,
or
6500
ppm
in
the
diet
for
24
months.

Melnick
et
al.
(
1998)
exposed
female
B6C3F
1
mice
(
10
animals/
group)
to
bromoform
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
Doses
of
bromoform
used
in
this
study
were
0
(
vehicle
only),
200,
or
500
mg/
kg­
day.
A
dose­
related
increase
in
absolute
liver
weight
and
liver
weight/
body
weight
ratio
was
noted
in
both
tested
doses.
Neither
serum
ALT
nor
serum
SDH
activity
was
significantly
elevated
at
either
dose
of
bromoform.
At
necropsy,
there
was
no
evidence
of
hepatocyte
hydropic
degeneration
in
animals
treated
with
either
dose.
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
Only
the
500
mg/
kg­
day
dose
resulted
in
marginally
significant
increase
in
hepatocyte
proliferation
as
measured
by
the
LI.
Using
the
Hill
equation
model,
these
authors
compared
the
dose
response
for
liver
toxicity
(
enzyme
and
LI
data)
and
tumorigenicity
(
using
data
from
previously
published
NTP
bioassays)
for
bromoform.
This
analysis
indicated
that
the
shape
of
the
dose
response
as
well
as
the
Hill
exponents
were
different
for
liver
toxicity
and
tumorigenicity.
The
authors
concluded
that
these
data
do
not
support
a
causal
relationship
between
liver
toxicity/
reparative
hyperplasia
and
tumor
development.

De
Angelo
et
al.
(
2002)
evaluated
the
ability
of
bromoform
administered
in
drinking
water
to
induce
aberrant
crypt
foci
(
ACF),
putative
early
preneoplastic
lesions,
in
the
colons
of
male
F344/
N
rats.
Groups
of
weanling
rats
(
6
animals/
group)
were
exposed
to
distilled
water,
0.25%
Alkamuls
EL­
620
®
,
or
1.1
g/
L
bromoform
in
0.25%
Alkamuls
EL­
620
for
13
weeks.
A
single
intraperitoneal
injection
of
30
mg/
kg
azoxymethane
(
AOM)
served
as
the
positive
control.
Body
weight
and
water
consumption
were
measured
twice
during
the
first
week
of
the
study
and
once
per
week
thereafter.
Colons
were
collected
at
study
termination,
fixed,
stained
with
0.2%
methylene
blue,
divided
into
three
equal
segments,
and
examined
for
ACF.
The
measured
concentration
of
bromoform
averaged
0.98
±
0.08
mg/
L
over
the
course
of
the
study.
When
adjusted
for
volatilization
and
adherence
to
glass,
the
corrected
concentration
was
0.77
mg/
L.
Water
consumption
was
significantly
reduced
(
30%)
in
the
bromoform
exposure
group
when
compared
to
the
deionized
water
control.
The
average
daily
dose
of
bromoform
was
76
mg/
kgday
as
calculated
by
the
study
authors.
The
average
terminal
body
weight
of
the
rats
exposed
to
bromoform
was
within
10%
of
the
control
values.
No
ACF
were
observed
in
colons
from
control
animals.
ACF
were
observed
in
four
of
six
colons
from
bromoform­
exposed
animals.
The
total
number
of
ACF
(
26)
and
number
of
aberrant
crypts
per
focus
(
3.71
±
0.36)
were
significantly
increased
relative
to
the
combined
deionized
water
and
vehicle
controls.
Fourteen
percent
of
the
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
101
observed
ACF
were
located
in
the
middle
segment
of
the
colon
and
86%
were
located
in
the
distal
(
rectal)
segment.
In
comparison,
807
ACF
and
4.95
±
0.25
crypts
per
focus
(
mean
and
standard
error)
were
observed
in
the
AOM
positive
control
group.
Eight
percent,
42%
and
50%
of
the
ACF
induced
by
AOM
were
located
in
the
proximal,
middle,
and
distal
segment
of
the
colon,
respectively.

H.
Other
Key
Health
Effects
1.
Immunotoxicity
a.
Bromodichloromethane
Munson
et
al.
(
1982)
administered
bromodichloromethane
by
gavage
to
CD­
1
male
and
female
mice
(
8­
12/
sex/
dose)
for
14
days
at
levels
of
0,
50,
125,
or
250
mg/
kg­
day.
Bromodichloromethane
appeared
to
affect
the
humoral
immune
system,
as
judged
by
decreased
antibody­
forming
(
ABF)
cells
in
serum
and
by
decreased
hemagglutination
titers.
These
changes
were
clearly
significant
(
p
<
0.05)
at
the
high
dose
in
both
males
and
females,
and
decreased
ABF
cells
were
also
noted
at
the
mid
dose
(
125
mg/
kg­
day)
in
females.
This
study
identified
a
NOAEL
of
50
mg/
kg­
day
and
a
LOAEL
of
125
mg/
kg­
day
for
bromodichloromethane
on
the
basis
of
decreased
immune
function
in
females
Additional
information
on
other
endpoints
measured
in
this
study
is
provided
in
Section
V.
B.
1.

French
et
al.
(
1999)
investigated
the
immunotoxicity
of
bromodichloromethane
in
a
series
of
four
experiments
conducted
in
mice
and
rats.
Immunotoxicity
in
mice
was
examined
following
exposure
via
ingestion
of
drinking
water
or
by
gavage.
The
immunological
parameters
examined
were
antibody
response
to
injected
sheep
red
blood
cells
and
T
and
B
lymphocyte
proliferation.
Mitogens
used
in
the
proliferation
assay
were
concanavalin
A
(
Con
A)
or
phyto­
hemagglutinin­
p
(
PHA)
for
T
cells
and
lipopolysaccharide
(
LPS)
for
B
cells.
Female
C57BL/
6
mice
(
6
animals
per
group)
were
treated
for
14
or
28
days
with
drinking
water
containing
0,
0.05,
0.25
or
0.5
g/
L
bromodichloromethane.
All
drinking
water
(
including
controls)
contained
0.25%
Emulphor
®
to
reduce
volatilization
of
bromodichloromethane.
Based
on
measured
water
consumption,
these
concentrations
were
estimated
by
the
authors
to
be
equivalent
to
0,
10,
37
or
62
mg/
kg­
day.
There
were
no
significant
differences
in
the
number
of
antibody
forming
cells,
antibody
production,
or
spleen
weights
in
any
treatment
group.
Likewise,
splenic
and
mesenteric
lymph
node
cell
proliferative
responses
to
T
and
B
cell
mitogens
were
similar
in
all
groups.
Continuation
of
this
study
for
an
additional
2
weeks
did
not
affect
any
measured
parameter.
These
data
identify
a
NOAEL
of
62
mg/
kg­
day
for
short­
term
exposure.

French
et
al.
(
1999)
conducted
a
second
experiment
in
which
female
C57BL/
6
mice
were
dosed
by
gavage
with
bromodichloromethane
in
10%
Emulphor
®
once
a
day
for
16
days.
Treatment
groups
(
6
animals
per
group)
included
controls
(
deionized
water
or10%
Emulphor
®
)
,
50,
125
or
250
mg/
kg­
day
bromodichloromethane.
As
in
the
previous
experiment,
there
were
no
differences
in
ABF
cells,
antibody
titers
or
mitogen­
induced
proliferation
in
any
treatment
groups.
A
decrease
in
spleen
weight
and
spleen­
to­
weight
ratio
was
observed
in
the
125
mg/
kg­
day
group
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
V
­
102
when
compared
to
the
Emulphor
®
control.
However,
spleen
weights
in
the
Emulphor
®
control
were
significantly
higher
than
those
in
the
deionized
water
control
group,
making
this
finding
difficult
to
interpret.

French
et
al.
(
1999)
investigated
the
immunotoxicity
of
bromodichloromethane
in
male
Fisher
344
rats
following
two
different
in
vivo
exposure
regimens:
ingestion
of
drinking
water
containing
bromodichloromethane
and
gavage.
The
immunological
parameters
examined
were
antibody
response
to
injected
sheep
red
blood
cells
and
T
and
B
lymphocyte
proliferation.
The
mitogens
used
in
the
proliferation
assay
were
concanavalin
A
(
Con
A)
or
phyto­
hemagglutin­
p
(
PHA)
for
T
cells
and
S.
typhimurium
mitogen
(
STM)
for
B
cells.
Six
rats
per
treatment
group
were
exposed
for
26
weeks
to
drinking
water
containing
0,
0.07
or
0.7
g/
L
bromodichloromethane
and
0.25%
Emulphor
®
.
Based
on
water
consumption
measurements,
these
concentrations
were
estimated
by
the
authors
to
be
equivalent
to
average
daily
doses
of
0,
5
or
49
mg/
kg­
day.
There
was
a
significant
suppression
of
Con
A­
stimulated
proliferation
of
spleen
cells
observed
in
the
49
mg/
kg­
day
dose
group.
No
effect
on
other
immunological
parameters
was
reported.
These
data
suggest
NOAEL
and
LOAEL
values
of
5
and
49
mg/
kg­
day,
respectively,
for
immunotoxic
effects.

French
et
al.
(
1999)
also
examined
the
effect
of
short­
term
exposure
to
relatively
large
doses
of
bromodichloromethane
on
immune
function.
Female
F344
rats
(
6
animals/
group)
received
gavage
doses
of
deionized
water,
10%
Emulphor
®
,
or
75,
150,
or
300
mg
bromodichloromethane/
kg
in
10%
Emulphor
®
for
5
days.
Surviving
high­
dose
animals
had
decreased
body,
spleen,
and
thymus
weights.
Con
A
and
PHA
responses
were
depressed
in
spleen
cells
isolated
from
high­
dose
animals.
Two
of
the
six
rats
in
the
300
mg/
kg­
day
group
died
during
the
exposure
period.
The
remaining
high­
dose
animals
had
significantly
decreased
body,
spleen
and
thymus
weights
compared
to
both
control
groups.
Thymus
weight,
but
not
spleen
or
body
weight,
was
also
decreased
in
the
150
mg/
kg­
day
group.
Con
A
responses
were
significantly
depressed
in
both
spleen
and
mesenteric
lymph
node
(
MLN)
cells
in
the
300
mg/
kgday
treatment
group.
All
three
(
75,
150
and
300
mg/
kg­
day)
dose
groups
exhibited
suppression
of
PHA
stimulated
MLN
cells
when
compared
to
the
vehicle
(
but
not
the
water)
controls.
This
discrepancy
was
due
to
the
fact
that
Emulphor
®
alone
significantly
elevated
the
proliferative
response
to
PHA
in
MLN
cells
relative
to
the
deionized
water
group.
In
contrast
to
the
T
cell
responses,
there
was
a
significant
increase
in
antibody
production
and
proliferative
responses
to
STM
(
B
cells)
from
spleen
cells
at
the
highest
dose
tested
(
300
mg/
kg­
day
dose
group).
These
data
suggest
a
marginal
NOAEL
of
150
mg/
kg­
day
and
a
LOAEL
of
300
mg/
kg­
day
for
acute
exposure
based
on
depression
of
immune
response.

b.
Dibromochloromethane
Munson
et
al.
(
1982)
administered
dibromochloromethane
by
gavage
to
CD­
1
male
and
female
mice
(
8
to
12/
sex/
dose)
for
14
days
at
levels
of
0,
50,
125,
or
250
mg/
kg­
day
and
evaluated
humoral
and
cell­
mediated
immune
system
functions.
Dibromochloromethane
appeared
to
affect
the
humoral
immune
system,
as
judged
by
decreased
antibody­
forming
(
ABF)
cells
in
serum
and
by
decreased
hemagglutination
titers.
These
changes
were
significant
Draft
­
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February
20,
2003
V
­
103
(
p
<
0.05)
at
the
high
dose
in
both
males
and
females.
Decreased
ABF
cells
were
also
noted
at
the
mid
dose
(
125
mg/
kg­
day)
in
females.
This
study
identified
a
NOAEL
of
50
mg/
kg­
day
and
a
LOAEL
of
125
mg/
kg­
day
for
dibromochloromethane
on
the
basis
of
decreased
immune
function
in
females.
Additional
information
on
this
study
is
provided
in
Section
V.
B.
2.

c.
Bromoform
Munson
et
al.
(
1982)
administered
bromoform
(
aqueous)
by
gavage
to
CD­
1
male
and
female
mice
(
6
to
12/
sex/
dose)
for
14
days
at
levels
of
0,
50,
125,
or
250
mg/
kg­
day.
Endpoints
evaluated
included
humoral
immune
system
function.
The
authors
judged
that
the
humoral
immune
system
was
not
significantly
affected
by
bromoform,
although
a
decrease
in
antibody
forming
(
ABF)
cells
was
reported
for
high­
dose
males.
These
data
suggest
a
NOAEL
of
250
mg/
kg­
day
for
effects
of
bromoform
on
the
immune
system.
Additional
information
on
this
study
is
provided
in
Section
V.
B.
3.

2.
Hormonal
disruption
No
studies
or
case
reports
were
identified
that
described
hormonal
disruption
by
dibromochloromethane
or
bromoform.

Bielmeier
et
al.
(
2001)
examined
the
effect
of
bromodichloromethane
on
serum
progesterone
and
luteinizing
hormone
levels
in
two
experiments
conducted
as
part
of
an
investigation
on
full
litter
resorption
(
FLR)
in
F344
rats
(
see
Section
V.
E.
1
for
additional
details).
In
the
first
experiment,
rats
(
7
to
10/
treatment
group)
were
administered
a
single
100
mg/
kg
dose
by
aqueous
gavage
on
gestation
day
8
or
9.
Hormone
levels
in
samples
of
tail
blood
were
determined
on
GD
9
through
12.
FLR
was
observed
in
0,
60
and
100%
of
the
control,
GD
8­
dosed,
and
GD
9­
dosed
animals,
respectively.
A
marked
reduction
in
progesterone
levels
was
noted
24
hours
after
dosing
in
all
rats
that
resorbed
their
litters
when
compared
to
controls
and
to
bromodichloromethane­
treated
animals
that
retained
their
litters.
The
mean
progesterone
levels
in
animals
dosed
on
GD
9
decreased
from
137.94
ng/
mL
±
11.44
ng/
mL
to
48.45
±
23.57
ng/
mL
within
24
hours
(
n
=
9).
For
animals
treated
on
GD
8,
the
mean
progesterone
level
24
hours
after
bromodichloromethane
treatment
was
67.01
±
16.22
ng/
mL
in
animals
that
resorbed
litters
(
n
=
6)
and
127.19
±
14.89
in
controls
(
n=
7).
The
resorbed
groups
had
reduced
progesterone
levels
comparable
to
the
progesterone
levels
in
non­
pregnant
animals
(
n
=
2)
when
assayed
three
days
after
compound
administration.
In
contrast
to
the
effect
noted
on
progesterone
levels,
administration
of
bromodichloromethane
had
no
apparent
effect
on
LH
when
measured
24
hours
after
dosing.
However,
elevated
LH
concentrations
were
observed
on
GD
11
to
12
in
animals
experiencing
resorption.
LH
levels
in
these
groups
increased
from
approximately
0.20
ng/
mL
on
GD
10
to
approximately
0.80
ng/
mL
on
GD
11
and
remained
elevated
through
GD
12.
In
contrast,
LH
levels
in
the
controls
decreased
from
0.31
to
0.14
ng/
mL
over
the
same
time
period.

Bielmeier
et
al.
(
2001)
performed
a
second
experiment
to
further
characterize
the
effect
of
bromodichloromethane
treatment
on
progesterone
and
LH
levels
in
pregnant
F344
rats.
The
rats
(
8­
11/
treatment
group)
were
dosed
with
0,
75,
or
100
mg/
kg
by
aqueous
gavage
on
GD
9.
Blood
Draft
­
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or
Quote
February
20,
2003
V
­
104
samples
were
collected
at
0,
6,
12,
and
24
hours
after
dosing.
The
incidence
of
FLR
was
0,
64%,
and
90%
in
the
0,
75,
and
100
mg/
kg
dose
groups,
respectively.
The
progesterone
levels
peaked
in
all
dose
groups
(
including
controls)
at
6
hours.
At
12
and
24
hours,
the
progesterone
levels
in
bromodichloromethane­
treated
animals
that
resorbed
their
litters
were
progressively
reduced.
Progesterone
levels
in
bromodichloromethane­
treated
animals
that
retained
their
litters
remained
comparable
to
levels
observed
in
the
control
group.
No
significant
differences
in
LH
concentration
were
noted
among
dose
groups
at
any
time
point.

3.
Structure­
Activity
Relationships
Although
the
mechanism
of
brominated
trihalomethane
toxicity
is
not
known
with
certainty,
there
is
abundant
evidence
to
indicate
that
adverse
effects
are
secondary
to
metabolism.
Bromine
is
generally
a
better
leaving
group
than
chlorine,
suggesting
that
bromine
substitution
could
potentially
influence
the
pathway
and
rate
of
trihalomethane
metabolism.
Multiple
studies
(
described
in
Section
III.
C)
indicate
that
metabolism
of
chloroform
and
the
brominated
trihalomethanes
can
occur
through
one
or
both
of
two
cytochrome
P450­
mediated
pathways:
reductive
metabolism
to
free
radical
intermediates
or
oxidative
metabolism
to
dihalocarbonyls
(
Figure
4­
1).
Although
comparative
data
are
limited,
there
is
some
evidence
to
indicate
that
chloroform
and
the
brominated
trihalomethanes
are
metabolized
to
a
different
extent
by
these
pathways.
Tomasi
et
al.
(
1985)
examined
the
reductive
metabolism
of
chloroform,
bromodichloromethane,
and
bromoform
in
rats
and
obtained
the
following
rank
order
for
free
radical
formation:
bromoform>
bromodichloromethane>
chloroform.
Wolf
et
al.
(
1977)
reported
that
bromoform
was
more
extensively
metabolized
under
anaerobic
conditions
in
vitro
than
was
chloroform.
Gao
and
Pegram
(
1992)
observed
that
binding
of
reactive
intermediates
to
rat
hepatic
microsomal
lipids
and
proteins
was
more
than
twice
as
high
for
bromodichloromethane
as
for
chloroform
when
assayed
under
anaerobic
conditions.
These
results
collectively
suggest
that
reductive
metabolism
may
be
a
more
important
metabolic
pathway
for
brominated
trihalomethanes
than
for
chloroform.
At
present,
this
apparent
difference
in
metabolism
has
not
been
linked
to
specific
differences
in
toxicity.

Two
mutagenicity
studies
provide
additional
information
on
structure­
activity
relationships
among
the
trihalomethanes.
Additional
details
of
these
studies
are
presented
in
Section
V.
F.
Examination
of
mutagenicity
in
a
strain
of
Salmonella
typhimurium
engineered
to
express
rat
theta­
class
glutathione­
S­
transferase
(
GST)
indicated
the
following
order
for
mutagenic
potency
(
number
of
revertants/
ppm)
of
the
brominated
trihalomethanes:
bromoform

dibromochloromethane>
bromodichloromethane
(
DeMarini
et
al.,
1997).
The
potency
of
the
first
two
compounds
was
several
times
greater
than
that
observed
for
bromodichloromethane.
Analysis
of
the
mutational
spectra
of
the
brominated
trihalomethanes
indicated
that
all
three
compounds
have
similar
mutational
spectra
(
predominately
GC

AT
transitions)
and
site
specificity
(
middle
C
of
a
CCC
sequence
in
target
DNA).
These
observations
suggest
that
a
common
reactive
intermediate
or
class
of
intermediates
is
likely
to
mediate
the
mutagenicity
of
these
compounds.
Draft
­
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Quote
February
20,
2003
V
­
105
In
the
second
study,
Pegram
et
al.(
1997)
compared
the
glutathione
S­
transferase­
mediated
mutagenicity
of
bromodichloromethane
and
chloroform
in
a
GST+
strain
of
S.
typhimurium
(
See
section
V.
F.
1).
Revertants
were
produced
in
a
dose­
related
manner
in
the
presence
of
low
as
well
as
high
concentrations
of
bromodichloromethane.
In
contrast,
chloroform
induced
a
doubling
of
the
number
of
revertants
only
at
high
concentrations.
This
result
provides
evidence
that
bromine
substitution
of
trihalomethanes
confers
the
capability
for
GST­
catalyzed
transformation
to
mutagenic
intermediates
at
low
substrate
concentrations.
These
data
further
suggest
that
chloroform
and
the
brominated
trihalomethanes
may
induce
adverse
effects
via
different
modes
of
action,
and
indicate
the
need
for
care
in
extrapolating
the
characteristics
of
chloroform
metabolism
and
toxicity
to
brominated
trihalomethanes.

I.
Summary
1.
Health
Effects
of
Acute
and
Short
Term
Exposure
of
Animals
Large
oral
doses
of
brominated
trihalomethanes
are
lethal
to
laboratory
animals.
Reported
acute
LD
50
values
range
from
450
to
969
mg/
kg
for
bromodichloromethane,
800
to
1,200
mg/
kg
for
dibromochloromethane,
and
1,388
to
1,550
mg/
kg
for
bromoform.
Acute
lethality
values
are
summarized
in
Table
V­
1.

Acute
oral
exposure
to
sublethal
doses
of
brominated
trihalomethanes
can
also
produce
effects
on
the
central
nervous
system,
liver,
kidney,
and
heart.
Acute
duration
studies
investigating
endpoints
other
than
death
are
summarized
in
Table
V­
2.
Ataxia,
anaesthesia,
and/
or
sedation
were
noted
in
mice
receiving
500
mg/
kg
bromodichloromethane,
500
mg/
kg
dibromochloromethane,
or
1,000
mg/
kg
bromoform.
Renal
tubule
degeneration,
necrosis,
and
elevated
levels
of
urinary
markers
of
renal
toxicity
have
been
observed
in
rats
receiving
200
to
400
mg/
kg
bromodichloromethane.
Elevated
levels
of
serum
markers
for
hepatotoxicity
and
have
been
observed
in
rats
at
doses
of
bromodichloromethane
ranging
from
approximately
82
to
400
mg/
kg­
day,
and
hepatocellular
degeneration
and
necrosis
were
observed
at
400
mg/
kg.
Effects
on
heart
contractility
were
reported
in
male
rats
at
doses
of
333
and
667
mg/
kg
dibromochloromethane.

Short
term
studies
of
brominated
trihalomethanes
are
summarized
in
Table
V­
3.
Short­
term
exposure
of
laboratory
animals
to
brominated
trihalomethanes
has
been
observed
to
cause
effects
on
the
liver
and
kidney.
Hepatic
effects,
including
organ
weight
changes,
elevated
serum
enzyme
levels,
and
histopathological
changes,
became
evident
in
mice
and/
or
rats
administered
38
to
250
mg/
kg­
day
bromodichloromethane,
147
to
500
mg/
kg­
day
dibromochloromethane,
or
187
to
289
mg/
kg­
day
bromoform
for
14
to
30
days.
Kidney
effects,
characterized
by
decreased
p­
aminohippurate
uptake,
histopathological
changes,
and
organ
weight
changes,
became
evident
in
mice
and/
or
rats
administered
148
to
300
mg/
kg­
day
bromodichloromethane,
147
to
500
mg/
kg­
day
dibromochloromethane,
or
289
mg/
kg­
day
bromoform
for
14
days.
Evidence
for
decreased
immune
function
was
noted
at
bromodichloromethane
or
dibromochloromethane
doses
of
125
mg/
kg­
day.
Studies
examining
Draft
­
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or
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February
20,
2003
V
­
106
strain
differences
in
response
to
short­
term
brominated
trihalomethane
exposure
have
not
been
reported.

2.
Health
Effects
of
Longer­
term
Exposure
of
Animals
Subchronic
studies
of
brominated
trihalomethanes
are
summarized
in
Table
V­
4.
The
predominant
effects
of
subchronic
oral
exposure
occur
in
the
liver
and
kidney.
The
effects
produced
in
these
two
organs
are
similar
in
nature
to
those
described
for
short­
term
exposures,
with
liver
appearing
to
be
the
most
sensitive
target
organ
for
dibromochloromethane
and
bromoform
exposure.
Histopathological
changes
in
the
liver
were
reported
in
mice
and/
or
rats
administered
200
mg/
kg­
day
bromodichloromethane,
50
to
250
mg/
kg­
day
dibromochloromethane
or
50
to
283
mg/
kg­
day
bromoform.
Histopathological
changes
in
the
kidney
were
reported
in
mice
and/
or
rats
administered
100
mg/
kg­
day
bromodichloromethane,
or
250
mg/
kgday
dibromochloromethane.
Studies
examining
strain
differences
in
response
to
subchronic
brominated
trihalomethane
exposure
have
not
been
reported.

Chronic
toxicity
studies
of
brominated
trihalomethanes
are
summarized
in
Table
V­
5.
As
observed
for
exposure
for
shorter
durations,
the
predominant
effects
of
chronic
oral
exposure
are
observed
in
the
liver
and
kidney.
Histopathological
signs
of
hepatic
toxicity
in
mice
and/
or
rats
became
evident
at
doses
of
6
to
50
mg/
kg­
day
for
bromodichloromethane,
40
to
50
mg/
kg­
day
for
dibromochloromethane,
and
90
to
152
mg/
kg­
day
for
bromoform.
Signs
of
bromodichloromethane­
induced
renal
toxicity
became
evident
in
mice
and
rats
treated
with
doses
of
25
and
50
mg/
kg­
day,
respectively.
Studies
examining
strain
differences
in
response
to
chronic
brominated
trihalomethane
exposure
have
not
been
reported.

3.
Reproductive
and
Developmental
Effects
Reproductive
and
developmental
studies
of
brominated
trihalomethanes
are
summarized
in
Table
V­
9.
Data
on
the
developmental
effects
of
brominated
trihalomethanes
suggest
that
these
chemicals
are
toxic
to
the
fetus
in
most
cases
only
at
doses
that
result
in
maternal
toxicity.
Signs
of
maternal
toxicity
(
decreased
body
weight,
body
weight
gain
and/
or
changes
in
organ
weight)
were
reported
in
rats
administered
bromodichloromethane
at
25
to
200
mg/
kg­
day
and
in
rabbits
administered
4.9
to
35.6
mg/
kg­
day.
Signs
of
maternal
toxicity
were
observed
in
rats
or
mice
administered
17
(
marginal)
to
200
mg/
kg­
day
dibromochloromethane
and
in
mice
administered
100
mg/
kg­
day
bromoform.
Maternal
toxicity
was
not
observed
in
female
rats
dosed
with
up
to
200
mg/
kg­
day
of
bromoform.
Several
well­
conducted
studies
on
the
developmental
toxicity
of
bromodichloromethane
gave
negative
results
at
doses
up
to
116
mg/
kg­
day
in
rats
and
76
mg/
kgday
in
rabbits
when
administered
in
drinking
water.
However,
in
other
studies,
slightly
decreased
numbers
of
ossification
sites
in
the
hindlimb
and
forelimb
were
observed
in
fetuses
of
rats
administered
45
mg/
kg­
day
in
the
drinking
water
on
gestation
days
6
to
21and
sternebral
aberrations
were
observed
in
the
offspring
of
rats
administered
200
mg/
kg­
day
by
gavage
in
corn
oil.
Reductions
in
mean
pup
weight
gain
and
pup
weight
were
observed
when
the
pups
were
administered
bromodichloromethane
in
the
drinking
water
at
concentrations
of
150
ppm
and
above
(
biologically
meaningful
estimates
of
intake
on
a
mg/
kg­
day
basis
could
not
be
calculated
Draft
­
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Not
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or
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20,
2003
V
­
107
for
this
study).
Full
litter
resorption
has
been
noted
in
F344
rats,
but
not
Sprague­
Dawley
rats,
treated
with
bromodichloromethane
at
doses
of
50
to
100
mg/
kg­
day
during
gestation
days
6
to
10.
Chronic
oral
exposure
to
bromodichloromethane
resulted
in
reduced
sperm
velocities
at
doses
of
39
mg/
kg­
day.
This
response
was
not
accompanied
by
histopathological
changes
in
any
reproductive
tissue
examined.
Adverse
clinical
signs
and
reduced
body
weight
and
body
weight
gain
were
observed
in
parental
generation
female
rats
and
F
1
male
and
female
rats
at
150
ppm
(
approximately
11.6
to
40.2
mg/
kg­
day)
in
a
two
generation
study
of
bromodichloromethane
administered
in
drinking
water.
In
the
same
study,
small
but
statistically
significant
delays
in
sexual
maturation
occurred
in
F
1
males
at
50
ppm
(
approximately
11.6
to
40.2
mg/
kg­
day)
and
F1
females
at
450
ppm
(
approximately
29.5
to
109
mg/
kg­
day).
These
delays
may
have
been
secondary
to
dehydration
caused
by
taste
aversion
to
bromodichloromethane
in
the
drinking
water.

Four
of
five
studies
on
the
reproductive
or
developmental
toxicity
of
dibromochloromethane
gave
negative
results
when
tested
at
doses
of
up
to
200
mg/
kg­
day.
In
the
fifth
study,
dibromochloromethane
administered
at
17
mg/
kg­
day
in
a
multigenerational
study
resulted
in
reduced
day
14
postnatal
in
one
of
two
F2
generation
litters.
At
171
mg/
kg­
day,
the
mid­
dose
in
the
study,
decreased
litter
size,
viability
index,
lactation
index,
and
postnatal
body
weight
were
observed
in
some
F1
and/
or
F2
generation.
The
developmental
and
reproductive
toxicity
of
bromoform
was
examined
in
two
studies.

Bromoform
administered
to
rats
at
100
mg/
kg­
day
in
corn
oil
by
gavage
resulted
in
a
significant
increase
in
sternebral
aberrations
in
the
apparent
absence
of
maternal
toxicity.
In
a
continuous
breeding
toxicity
protocol,
gavage
doses
of
200
mg/
kg­
day
in
corn
oil
resulted
decreased
postnatal
survival,
organ
weight
changes,
and
liver
histopathology
in
F1
mice
of
both
sexes.
No
effects
on
fertility
or
other
reproductive
endpoints
were
noted.

4.
Mutagenicity
and
Genotoxicity
In
vitro
and
in
vivo
studies
of
the
mutagenic
and
genotoxic
potential
of
bromodichloromethane,
dibromochloromethane,
and
bromoform
have
yielded
mixed
results.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid),
and
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
Overall,
a
majority
of
studies
yielded
more
positive
results
for
bromoform
and
bromodichloromethane.
The
genotoxicity
and
mutagenicity
data
for
dibromochloromethane
are
variable.
Recent
studies
in
strains
of
Salmonella
engineered
to
contain
rat
theta­
class
glutathione
S­
transferase
suggest
that
mutagenicity
of
the
brominated
trihalomethanes
may
be
mediated
by
glutathione
conjugation.

5.
Carcinogenicity
Studies
in
Animals
The
carcinogenic
potential
of
individual
brominated
trihalomethanes
administered
in
oil
has
been
investigated
in
chronic
oral
exposure
studies
in
mice
and
rats.
Ingestion
of
Draft
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20,
2003
V
­
108
bromodichloromethane
caused
liver
tumors
in
female
mice,
renal
tumors
in
male
mice
and
in
male
and
female
rats,
and
tumors
of
the
large
intestine
in
male
and
female
rats.
Ingestion
of
dibromochloromethane
caused
liver
tumors
in
male
and
female
mice,
and
ingestion
of
bromoform
caused
intestinal
tumors
in
male
and
female
rats.
Comparison
of
dose
response
data
for
hepatotoxicity,
cell
proliferation,
and
tumorigenesis
in
female
mice
suggests
that
the
hepatic
carcinogenicity
of
brominated
trihalomethanes
is
not
a
simple
consequence
of
cytotoxicity
and
regenerative
cell
proliferation
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20,
2003
V
­
109
6.
Other
Key
effects
The
immunotoxicity
of
brominated
trihalomethanes
has
been
investigated
in
mice
and
rats.
Short­
term
bromodichloromethane
exposure
resulted
in
decreased
antibody
forming
cells
in
serum,
decreased
hemagglutinin
titers,
and/
or
suppression
of
Con
A­
stimulated
proliferation
of
spleen
cells
at
doses
of
125
to
250
mg/
kg­
day.

No
evidence
has
been
reported
for
hormonal
effects
following
exposure
to
dibromochloromethane
or
bromoform.
Studies
in
pregnant
F344
rats
detected
decreased
levels
of
progesterone
in
animals
administered
75
or
100
mg/
kg
bromodichloromethane
by
aqueous
gavage
on
gestation
day
8
or
9.
Increased
levels
of
luteinizing
hormone
were
observed
two
to
three
days
after
dose
administration.
Disruption
of
luteal
responsiveness
to
luteinizing
hormone
has
been
proposed
as
a
possible
mode
of
action
by
which
bromodichloromethane
elicits
full
litter
resorption
in
F344
rats.

Limited
structure­
activity
data
for
brominated
trihalomethanes
and
chloroform
suggest
that
bromination
may
influence
the
proportion
of
compound
metabolized
via
the
oxidative
and
reductive
pathways,
with
brominated
compounds
being
more
extensively
metabolized
by
the
reductive
pathway.
Additional
evidence
suggests
that
a
GSH­
mediated
pathway
may
play
an
important
role
in
metabolism
of
brominated
trihalomethanes,
but
not
chloroform.
These
data
raise
the
possibility
that
brominated
trihalomethanes
may
induce
some
adverse
effects
via
different
modes
of
action
than
chloroform.
Draft
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20,
2003
VI
­
1
VI.
HEALTH
EFFECTS
IN
HUMANS
A.
Clinical
Case
Studies
1.
Bromodichloromethane
No
clinical
reports
or
short
term
studies
were
located
on
the
effects
in
humans
from
ingestion
of
bromodichloromethane.

2.
Dibromochloromethane
No
clinical
case
reports
or
short
term
studies
were
located
on
the
effects
in
humans
from
ingestion
of
dibromochloromethane.

3.
Bromoform
In
the
past,
bromoform
was
used
as
a
sedative
for
children
with
whooping
cough.
Typical
doses
were
approximately
one
drop
(
about
180
mg),
given
three
to
six
times/
day
(
Burton­
Fanning,
1901).
This
dosing
usually
resulted
in
mild
sedation
in
children,
although
a
few
rare
instances
of
death
or
near­
death
were
reported
(
e.
g.,
Dwelle,
1903;
Benson,
1907).
These
cases
were
believed
to
be
due
to
accidental
overdoses.
Based
on
these
clinical
observations,
the
estimated
lethal
dose
for
a
10­
to
20­
kg
child
is
approximately
300
mg/
kg,
and
the
LOAEL
for
mild
sedation
is
approximately
54
mg/
kg­
day.

B.
Epidemiological
Studies
Multiple
epidemiological
studies
have
investigated
the
relationship
between
exposure
to
disinfection
by­
products
in
chlorinated
drinking
water
and
adverse
health
effects.
These
studies
fall
into
two
basic
categories:
studies
of
association
with
cancer
(
Table
VI­
1)
and
studies
of
association
with
adverse
pregnancy
or
birth
outcomes
or
alteration
of
menstrual
function
(
Table
VI­
2).
Because
the
purpose
of
this
document
is
to
isolate
the
health
effects
of
individual
brominated
trihalomethanes,
a
detailed
examination
of
all
available
studies
on
disinfection
byproducts
is
beyond
the
scope
of
this
report.
Epidemiologic
studies
published
prior
to
1994
are
discussed
in
greater
detail
in
the
Drinking
Water
Criteria
on
Chlorine
(
U.
S.
EPA,
1994a).
A
number
of
recent
publications
have
reviewed
the
association
between
chlorination
disinfection
byproducts
and
cancer
and
adverse
reproductive
or
developmental
outcomes
(
e.
g.,
Reif
et
al.,
1996;
Mills
et
al.,
1998;
Nieuwenhuijsen
et
al.,
2000;
Bove
et
al.,
2002).

A
subset
of
epidemiologic
studies
has
examined
possible
associations
between
exposure
to
bromodichloromethane
and
adverse
reproductive
outcomes.
The
relationship
between
exposure
to
brominated
trihalomethanes
and
alterations
in
menstrual
function
has
also
been
investigated.
These
studies
are
described
in
greater
detail
in
the
sections
below.
Draft
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20,
2003
VI
­
2
Table
VI­
1
Epidemiological
Studies
Investigating
an
Association
Between
Chlorinated
Drinking
Water
and
Cancer
Reference
Study
Description
Observation
Alavanja
et
al.
(
1978)
Case
control
study
in
seven
New
York
State
counties.
Greater
risk
of
gastrointestinal
and
urinary
tract
cancer
mortality,
both
sexes,
in
chlorinated
water
areas
of
the
counties.

Cantor
et
al.
(
1978)
Ecological
study
using
age­
standardized
cancer
mortality
rates,
1968­
1971;
and
halomethane
levels
from
U.
S.
EPA
surveys.
Strongest
correlation
between
brominecontaining
trihalomethanes
and
bladder
cancer.

Struba
(
1979)
Case­
control
study
of
mortality
in
North
Carolina,
1975­
1978.
Small
but
significant
odds
ratios
for
rectum,
colon
and
bladder
cancers
in
rural
areas
but
not
in
urban
areas.

Brenniman
et
al.
(
1980)
Case­
control
study
in
70
Illinois
communities,
1973­
1976.
Questionnaires
sent
to
water
treatment
plants
to
verify
1963
inventory
data
on
chlorine
levels.
Statistically
significant
relative
risks
of
cancer
of
gall
bladder,
large
intestine,
and
total
gastrointestinal
and
urinary
tract
in
females
served
by
systems
with
chlorinated
versus
nonchlorinated
ground
water.
Due
to
many
uncontrolled
confounding
factors,
authors
concluded
that
chlorination
was
not
a
major
factor
in
the
etiology
of
gastrointestinal
and
urinary
tract
cancers.

Gottlieb
et
al.
(
1981)
Case­
control
study
using
mortality
data
in
Louisiana
and
estimations
of
exposure.
Rectal
cancer
significantly
elevated
with
respect
to
surface
or
Mississippi
River
water
consumption.

Young
et
al.
(
1981)
Case­
control
study
in
Wisconsin,
1972­
1977.
Questionnaires
sent
to
waterworks
superintendents
on
chlorine
content.
Colon
cancer
showed
significant
(
p<
0.05)
association
with
chlorine
intake
in
all
three
dosage
categories.

Cragle
et
al.
(
1985)
Case­
control
study
using
colon
cancer
cases
from
seven
hospitals
in
North
Carolina.
Consumption
of
chlorinated
water
strongly
associated
with
colon
cancer,
above
age
60.

Young
et
al.
(
1987)
Case­
control
study
of
colon
cancer
cases
in
Wisconsin.
Water
consumption
was
determined
by
interview,
and
chloroform
levels
by
historical
records
and
measurement.
No
association
found
between
trihalomethane
exposure
and
colon
cancer
incidence.
Table
VI­
1
(
cont.)

Reference
Study
Description
Observation
Draft
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or
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20,
2003
VI
­
3
Morris
et
al.
(
1992)
Meta­
analysis
of
nine
case­
control
studies
and
one
cohort
study
analyzing
cancer
and
consumption
of
chlorinated
water
or
water
containing
high
chloroform
levels.
Statistically
significant
relative
risk
of
rectal
cancer
and
bladder
cancer
in
exposed
groups.
No
colon
cancer.

McGheehin
et
al.
(
1993)
Population­
based
case­
control
study
Association
between
bladder
cancer
risk
and
exposure
to
chlorinated
water
and
trihalomethanes.

King
and
Marret
(
1996)
Case­
control
study
conducted
by
Health
Canada
Increased
risk
of
bladder
cancer
associated
with
total
trihalomethane
exposure.

Hildesheim
et
al.
(
1997)
Population­
based
case­
control
study
of
colon
and
rectal
cancer
risk.
Iowa,
1986­
1989.
Rectal
cancer
risk
associated
with
duration
of
chlorinated
water
use.
No
association
of
colon
cancer
risk
with
duration
of
chlorinated
water
use
or
trihalomethane
estimates.

Cantor
et
al.
(
1998)
Population­
based
case­
control
study
of
bladder
cancer
risk.
Iowa,
1986­
1989.
Positive
findings
for
risk
restricted
to
men
and
to
current
or
former
smokers.
In
men,
smoking
and
exposure
to
chlorinated
water
enhanced
the
risk
of
bladder
cancer.

Marcus
et
al.
(
1998)
Ecologic
study
of
association
between
TTHM
in
71
North
Carolina
public
water
supplies
and
incidence
of
histologically
confirmed
female
invasive
breast
cancer
obtained
from
cancer
registry
data.
TTHM
levels
not
associated
with
breast
cancer
risk
when
adjusted
for
potential
confounding
factors.
Data
were
consistent
with
TTHMs
being
unrelated
or
weakly
related
to
breast
cancer
risk.

Table
VI­
2
Epidemiological
Studies
Investigating
an
Association
Between
Chlorinated
Drinking
Water
and
Adverse
Pregnancy
Outcomes
or
Altered
Menstrual
Function
Table
VI­
2
(
cont.)

Draft
­
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Cite
or
Quote
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20,
2003
VI
­
4
Reference
Study
Description
Observation
Aschengrau
et
al.
1989
Hospital­
based
case­
control
study
of
spontaneous
abortion
and
multiple
water
quality
parameters
in
Boston,
MA
area.
After
adjustment
for
potential
confounders
and
chemical
constituents,
frequency
of
spontaneous
abortion
was
increased
for
consumption
of
surface
water
when
compared
to
use
of
mixed
surface
and
ground
water
(
OR
2.2,
95%
C.
I.
1.3
­
3.6)
The
association
between
surfance
water
and
increased
risk
of
spontaneous
abortion
was
not
confirmed
by
a
comparison
of
chlorinated
vs.
chloraminated
surface
water.
Chloraminated
water
was
used
as
a
surrogate
for
low
exposure
to
disinfection
by­
products.

Aschengrau
et
al.
1993
Case­
control
study
of
drinking
water
quality
and
occurrence
of
late
adverse
effects
among
women
who
delivered
infants
during
August
1977
­
March
1980
in
Massachusetts
After
adjustment
for
confounding,
frequency
of
stillbirths
was
increased
for
women
exposed
to
chlorinated
surface
water
(
OR
2.6,
95%
CI
0.9­
7.5).

Nuckols
et
al.
(
1995)
Cross­
sectional
study
in
Colorado
of
populations
drinking
chlorinated
and
chloraminated
water
No
statistically
significant
effects
of
exposure,
although
odds
ratio
was
elevated
for
risk
of
low
birth
weight
infants.

Bove
et
al.
(
1995)
Cross­
sectional
study
in
New
Jersey
Relationship
between
total
trihalomethane
levels
and
"
small
for
gestational
age."

Savitz
et
al.
(
1995)
Population
based
case­
control
study
in
North
Carolina
Statistically
significant
association
of
miscarriage
with
increasing
concentration
of
TTHM
and
with
the
highest
sextile
of
exposure(
OR=
2.8,
95%
C.
I.
1.1,
2.7),
but
no
relationship
with
ingested
dose
or
water
source.
Small
increase
in
risk
of
low
birth
rate.

Gallagher
et
al.
(
1998)
Retrospective
cohort
study
of
relationship
between
THM
exposure
during
third
trimester
of
pregnancy
and
low
birthweight,
low
term
birth
weight,
and
preterm
delivery.
Colorado
birth
certificate
data
matched
to
historical
water
data
based
on
census
block
groups.
1990­
1993.
Possible
association
of
trihalomethane
concentration
in
tap
water
at
maternal
residence
during
third
trimester
and
risk
of
term
low
birth
weight
deliveries.
Little
association
with
preterm
delivery.
Weak
association
with
low
birth
weight.
Table
VI­
2
(
cont.)

Reference
Study
Description
Observation
Draft
­
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Not
Cite
or
Quote
February
20,
2003
VI
­
5
Waller
et
al.
(
1998)
Prospective
study
of
association
between
total
and
individual
THM
exposure
and
spontaneous
abortion.
Concurrent
THM
data
obtained
from
public
water
supplies.
Women
who
drank

5
glasses/
day
of
cold
tap
water
containing

75
µ
g/
L
TTHMs
had
an
adjusted
odds
ratio
of
1.8
for
spontaneous
abortion.
Of
individual
THMs,
only
consumption
of

5
glasses
of
water
containing

18
µ
g/
L
bromodichloromethane
(
or
a
compound
co­
occurring
with
bromodichloromethane)
was
associated
with
spontaneous
abortion.

Klotz
and
Pyrch
(
1998;
1999)
Case­
control
study
of
association
between
drinking
water
contaminants
(
including
disinfection
byproducts)
and
neural
tube
defects.
Births
with
neural
tube
defects
reported
to
New
Jersey's
Birth
Defects
Registry
in
1993
and1994
were
matched
against
control
births
chosen
randomly
from
across
the
State.
Elevated
odds
ratios,
generally
between
1.5
and
2.1,
for
the
association
of
neural
tube
defects
with
total
THMs
(
TTHMs).
The
only
statistically
significant
results
were
seen
when
the
analysis
was
isolated
to
those
subjects
with
the
highest
THM
exposures
(
greater
than
40
ppb)
and
was
limited
to
those
subjects
with
neural
tube
defects
in
which
there
were
no
other
malformations
(
OR
=
2.1,
95%
CI
=
1.1
 
4.0).

Dodds
et
al.
(
1999)
Retrospective
cohort
study
in
Nova
Scotia
women
with
singleton
births,
1988­
1995.
Little
association
between
TTHM
level
and
fetal
weight­
or
gestational
agerelated
outcomes.
Elevated
relative
risk
for
stillbirths
for
exposure
to
>
100
µ
g/
L
TTHM
levels
during
pregnancy.
Little
evidence
for
increased
prevalence
or
dose­
response
for
congenital
abnormalities
with
possible
exception
of
chromosome
aberrations
for
exposure
>
100
µ
g/
L.

Magnus
et
al.
(
1999)
Ecologic
study
in
Norway
of
chlorinated
water
consumption
and
birth
defects
observed
in
births
during
period
1993­
1995.
1994
data
on
water
quality
and
disinfection
practice.
Water
color
used
as
an
indicator
for
natural
organic
matter
content.
Among
141,077
births,
1.8%
had
birth
defects.
Adjusted
odds
ratios
(
high
color,
chlorination
vs.
low
color,
no
chlorination)
of
1.14
(
0.99­
1.31)
for
any
malformation;
1.26
(
0.61­
2.62)
for
neural
tube
defects;
and
1.9
(
1.10­
3.57)
for
urinary
tract
defects.

Yang
et
al.
(
2000)
Study
in
Taiwan
of
association
between
chlorination
of
drinking
water
and
low
birth
weight.
Examination
of
18,025
births
showed
no
association
between
consumption
of
chlorinated
drinking
water
and
low
birth
weight.
Table
VI­
2
(
cont.)

Reference
Study
Description
Observation
Draft
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or
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20,
2003
VI
­
6
King
et
al.
(
2000)
Population­
based
retrospective
cohort
study
in
Nova
Scotia,
Canada
to
examine
the
relationship
between
TTHM
or
individual
THMs
and
risk
for
stillbirth
of
fetuses
greater
than
500
grams.
Study
cohort
assembled
from
a
perinatal
database
and
consisted
of
49,756
singleton
births
that
occurred
between
1988
and
1995.
Risk
doubled
for
women
exposed
to
a
bromodichloromethane
level

20
µ
g/
L
when
compared
to
women
consuming
concentrations
of
less
than
5
µ
g/
L
(
relative
risk
=
1.98,
95%
confidence
interval
of
1.23
­
3.49).
When
categories
of
stillbirth
(
unexplained
deaths
and
asphyxia­
related
deaths)
were
examined,
relative
risk
estimates
for
asphyxia­
related
deaths
increased
by
32%
for
each
10
µ
g/
L
increase
in
bromodichloromethane
concentration.

Dodds
and
King
(
2001)
Retrospective
cohort
study
conducted
using
data
from
a
population­
based
perinatal
database
in
Nova
Scotia,
Canada
and
routine
water
monitoring
data.
The
cohort
consisted
of
women
who
had
a
singleton
birth
in
Nova
Scotia
between
1988
and
1995
and
who
lived
in
an
area
with
a
municipal
water
supply.
Exposure
to
bromodichloromethane
at
concentrations
of
20
µ
g/
L
and
over
was
associated
with
increased
risk
of
neural
tube
defects
(
adjusted
relative
risk
=
2.5;
95%
confidence
interval
1.2
to
5.1)
and
decreased
risk
of
cardiovascular
anomalies
(
adjusted
relative
risk
=
0.3;
95%
confidence
interval
0.2
to
0.7).
No
association
observed
for
bromodichloromethane
and
cleft
defects.

Waller
et
al.
(
2001)
Reanalysis
of
total
trihalomethane
exposure
data
reported
in
Waller
et
al.
(
1998).
The
study
authors
reported
no
apparent
advantage
in
using
a
closest­
site
(
vs.
utility­
wide)
measurement
approach
for
estimation
of
exposure
to
total
trihalomethanes.

Windham
et
al.
(
2003)
Prospective
study
of
association
between
total
and
individual
THM
exposure
and
menstrual
cycle
function.
Concurrent
THM
data
obtained
from
public
water
supplies.
Exposure
to
dibromochloromethane
and
sum
of
brominated
trihalomethanes
was
associated
with
a
reductions
in
length
of
the
menstrual
cycle
and
follicular
phase
of
the
menstrual
cycle,
suggesting
possible
effects
on
ovarian
function.
Concentrations
of

20
µ
g/
L
for
dibromochloromethane
and

45
µ
g/
L
for
total
brominated
trihalomethanes
were
associated
with
reductions
in
cycle
and
follicular
phase
lengths
of
approximately
one
day.
No
effect
was
noted
on
length
of
luteal
phase
or
duration
of
menses.
Draft
­
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or
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February
20,
2003
VI
­
7
In
their
assessment
of
available
data
on
the
available
data
for
reproductive
and
developmental
effects
of
disinfection
byproducts,
Reif
et
al.
(
1996)
stress
that
interpretation
of
epidemiologic
findings
for
these
contaminants
are
potentially
complicated
by
unmeasured
confounding
variables
and
misclassification
errors.
Smoking,
socioeconomic
status,
alcohol
consumption,
other
environmental
exposures,
and
reproductive
history
are
examples
of
confounding
variables
that
have
the
potential
to
bias
estimates
of
risk
in
studies
of
disinfection
byproducts
if
not
measured.
Misclassification
errors
can
arise
from
failure
to
account
for
spatial
and
temporal
variability
in
contaminant
measurement
s,
migration
of
study
participants,
incorrect
assumptions
related
to
water
source
or
use,
and
use
of
water
treatment
data
as
a
surrogate
for
tap
water
concentrations.
These
factors
may
result
in
under­
or
over­
classification
of
health
risks
associated
with
the
consumption
of
disinfected
water.
Of
greatest
concern
are
variables
or
errors
which
might
lead
to
underestimation
of
the
true
public
health
risks
associated
with
exposure
to
tap
water
containing
brominated
trihalomethanes.
The
positive
findings
in
studies
of
brominated
trihalomethanes
thus
form
a
foundation
for
further
studies,
but
should
be
interpreted
cautiously.

1.
Bromodichloromethane
Kramer
et
al.
(
1992)
conducted
a
population­
based
case­
control
analysis
to
determine
if
exposure
to
trihalomethanes
in
drinking
water
is
associated
with
low
birthweight,
prematurity,
or
intrauterine
growth
retardation
(
lower
than
the
5th
percentile
of
weight
for
gestational
age).
A
separate
analysis
was
conducted
for
each
endpoint,
using
five
randomly
selected
controls
for
each
affected
newborn.
Data
were
collected
from
Iowa
birth
certificates
from
January
1,
1989,
to
June
30,
1990;
the
study
population
was
restricted
to
residents
of
small
towns
where
all
of
the
drinking
water
was
derived
from
a
single
source
(
surface
water,
shallow
wells,
or
deep
wells).
Exposure
data
were
based
on
a
1987
municipal
water
survey;
birth
certificate
data
from
1987
were
not
used
because
data
on
maternal
smoking
status
first
became
available
in
1989.
The
study
authors
adjusted
for
maternal
age,
number
of
previous
children,
marital
status,
education,
adequacy
of
prenatal
care,
and
maternal
smoking.
An
association
was
observed
between
exposure
to
at
least
10
µ
g/
L
bromodichloromethane
and
intrauterine
growth
retardation
(
odds
ratio
=
1.7).
However,
the
confidence
interval
in
these
cases
included
1,
indicating
that
the
increases
were
not
statistically
significant.

Waller
et
al.
(
1998)
conducted
a
prospective
study
in
pregnant
women
to
examine
the
association
between
trihalomethanes
in
drinking
water
and
spontaneous
abortion
(
pregnancy
loss
at
20
or
less
completed
weeks
of
gestation).
The
study
participants
were
recruited
from
three
facilities
of
a
large
managed
health
care
organization
which
were
located
in
regions
of
California
that
primarily
received
either
mixed,
surface,
or
groundwater.
Recruitment
occurred
when
the
women
scheduled
their
first
prenatal
exam
after
confirmation
of
pregnancy.
A
group
of
5,342
subjects
completed
a
telephone
interview
that
obtained
information
on
demographics,
previous
pregnancy
history,
employment
status,
consumption
of
tap
and
bottled
water,
substance
use
(
alcohol,
tobacco,
and
caffeine),
and
other
factors.
At
the
time
of
enrollment
in
the
study,
each
woman
was
at
least
18
years
of
age,
at
13
or
less
weeks
of
gestation,
spoke
English
or
Spanish,
and
could
provide
with
certainty
the
date
of
her
last
menstrual
period.
Following
adjustment
for
Draft
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February
20,
2003
VI
­
8
elective
termination
of
pregnancy,
ectopic
or
molar
pregnancies,
and
multiple
gestations,
a
total
of
5,144
pregnancies
remained
for
analysis.

Waller
et
al.
(
1998)
quantified
exposure
to
trihalomethanes
by
estimating
the
subject's
daily
tap
water
intake
at
8
weeks
gestation.
Concentration
of
total
trihalomethanes
and
any
available
data
on
individual
trihalomethanes
were
obtained
directly
from
the
utility
supplying
drinking
water
to
a
subject's
address
or
zip
code.
Total
trihalomethane
levels
were
calculated
by
averaging
all
measurements
taken
by
the
utility
supplying
a
participant's
home.
Each
participant
was
assigned
a
personal
exposure
classification
(
high
or
low)
to
total
trihalomethanes
(
TTHM)
and
individual
trihalomethanes
(
THM)
based
on
the
following
criteria.
A
high
personal
exposure
to
TTHM
was
defined
as
drinking
5
or
more
glasses
of
cold
tap
water
per
day
and
having
a
TTHM
level
of
75
µ
g/
L
or
higher.
Low
personal
exposure
to
TTHM
was
defined
as
either
1)
drinking
less
than
5
glasses
of
cold
tap
water
per
day,
2)
having
a
TTHM
level
of
less
than
75
µ
g/
L,
or
3)
receiving
water
from
a
utility
that
provided
95%
or
greater
groundwater.
Personal
exposures
to
the
individual
THMs
(
bromoform,
bromodichloromethane
and
dibromochloromethane)
were
defined
in
a
similar
manner,
with
a
high
personal
exposure
being
defined
as
drinking
5
or
more
glasses
of
cold
tap
water
per
day
with
an
individual
brominated
THM
level
of
16
µ
g/
L
or
higher
for
bromoform,
18
µ
g/
L
or
higher
for
bromodichloromethane,
or
31
µ
g/
L
or
higher
for
dibromochloromethane.
Low
personal
exposures
to
the
individual
THMs
were
defined
as
either
1)
drinking
less
than
5
glasses
of
cold
tap
water
per
day,
2)
having
an
individual
THM
level
below
the
cutoff,
or
3)
having
a
TTHM
level
less
than
72
µ
g/
L
if
individual
THM
levels
were
not
reported.

The
authors
found
a
modest
association
between
consumption
of
trihalomethanecontaining
water
and
incidence
of
spontaneous
abortion.
Increased
risk
of
spontaneous
abortion
was
noted
starting
at
approximately
75
µ
g/
L.
The
adjusted
odds
ratio
(
OR)
for
women
who
drank
5
or
more
glasses
of
cold
tap
water
per
day
containing
an
average
TTHM
level
of
75
µ
g/
L
or
higher
during
their
first
trimester
was
1.8
(
95%
confidence
interval
(
C.
I.)
1.1,
3.0).
An
estimated
18.4%
of
the
study
participants
were
exposed
at
or
above
this
level.
Of
the
four
individual
THMs,
only
high
bromodichloromethane
exposure
(
or
exposure
to
another
compound
closely
associated
with
bromodichloromethane)
was
associated
with
spontaneous
abortion
alone
(
adjusted
OR
=
2.0,
95%
C.
I.
1.2,
3.5)
and
after
adjustment
for
other
THMs
(
adjusted
OR
=
3.0,
C.
I.
1.4,
6.6).
Waller
et
al.
(
1998)
noted
that
there
was
no
additive
or
other
effect
from
showering
or
swimming.
Therefore,
no
adjustment
was
required
for
these
variables.
Misclassification
of
exposure
was
identified
as
the
primary
limitation
of
this
study.
Concentration
levels
for
most
subjects
were
based
on
test
results
for
a
single
day,
and
thus
do
not
reflect
potential
variation
in
trihalomethane
levels
over
time.
In
addition,
the
exposure
to
THMs
from
sources
other
than
ingestion
could
not
be
fully
characterized.

Because
exposure
misclassification
appeared
to
be
a
limitation
of
the
Waller
et
al.
(
1998)
study,
Waller
et
al.
(
2001)
reported
a
reanalysis
of
exposure
data
from
that
study.
The
objective
of
the
new
analysis
was
to
examine
how
use
of
alternative
methods
for
estimation
of
exposure
would
affect
associations
between
TTHM
exposure
and
risk
of
spontaneous
abortion.
This
reanalysis
did
not
address
dose­
response
relationships
between
individual
brominated
Draft
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February
20,
2003
VI
­
9
trihalomethanes
and
occurrence
of
spontaneous
abortion.
Two
exposure
analyses
were
tested.
The
first
method
used
the
utility­
wide
average
concentration
(
the
metric
used
in
Waller
et
al.,
1998).
The
second
method
used
THM
measurements
taken
from
the
water
system
sampling
site
nearest
the
subject's
home.
For
each
method,
the
authors
performed
1)
an
unweighted
analysis;
2)
an
analysis
weighted
by
a
factor
intended
to
reduce
exposure
misclassification;
and
3)
an
analysis
within
a
subset
of
the
cohort
that
possibly
had
less
exposure
misclassification
than
the
entire
cohort.
The
weighted
and
subset
analyses
were
performed
in
an
effort
to
reduce
exposure
misclassification.
The
utility­
wide
average
method
estimated
the
concentration
of
total
trihalomethanes
by
averaging
all
distribution
measurements
taken
by
the
subject's
utility
during
the
first
trimester
of
pregnancy.
In
contrast
to
the
method
used
in
Waller
et
al.
(
1998)
the
time
interval
was
not
expanded
in
order
to
capture
a
measurement
and
thus
reduce
missing
data
in
cases
where
no
measurements
were
available
in
the
first
trimester.
The
closest­
site
method
took
the
average
of
all
measurements
taken
during
the
first
trimester
of
pregnancy
at
the
water
distribution
site
nearest
to
the
subject's
home.
For
the
weighted
analyses,
the
utility­
wide
approach
used
the
variance
of
the
utility­
wide
average
as
a
proxy
for
accuracy.
The
closest­
site
approach
used
TTHM
measurements
taken
from
the
water
system
sampling
site
nearest
the
subject's
home
and
adjusted
for
distance
between
the
subject's
home
and
the
sampling
site.
For
the
subset
approach,
analyses
were
restricted
to
groups
of
women
for
whom
the
exposure
assessment
was
likely
to
be
more
accurate.
Subset
analyses
using
the
utility­
wide
average
TTHM
concentration
included
women
whose
utility
measurements
were
all
within
20
µ
g/
L
of
each
other
and
women
served
by
groundwater
utilities.
Subset
analyses
using
the
closest­
site
average
concentrations
used
women
who
lived
within
0.5
miles
of
the
utility
sampling
site
and
all
women
served
by
groundwater
utilities.
An
ingestion
metric
was
calculated
using
individual
daily
cold
tap
water
intake
at
eight
weeks
gestation
as
determined
in
Waller
et
al.
(
1998).
A
categorical
ingestion
exposure
metric
was
created
using
the
first
trimester
THM
concentration
dichotomized
at
75
µ
g/
L
and
cold
tap
water
ingestion
dichotomized
at
5
glasses
per
day.
Ingestion
was
also
estimated
by
multiplying
the
TTHM
concentration
by
cold
tap
water
consumption.
A
metric
to
capture
exposure
to
trihalomethanes
during
showering
was
created
by
multiplying
THM
concentration
by
typical
shower
duration
and
the
frequency
of
showering.

Use
of
the
utility­
wide
approach
generally
resulted
in
odds
ratios
equivalent
to
or
slightly
higher
than
the
closest­
site
approach.
Odds
ratios
obtained
using
the
utility­
wide
average
method
for
estimating
TTHM
(
but
not
the
closest­
site
method)
became
progressively
stronger
in
the
weighted
and
subset
analyses.
The
study
authors
reported
a
positive,
monotonic
dose­
response
relationship
between
spontaneous
abortion
rate
and
an
exposure
metric
incorporating
TTHM
and
personal
ingestion.
Rates
of
spontaneous
abortion
obtained
using
this
approach
ranged
from
8.3%
to
13.7%
(
unweighted);
7.9%
to
16.6
%
(
variance
weighted);
and
6.6%
to
16.6%
(
lowvariance
subset).
The
study
authors
noted
that
a
major
limitation
of
this
reanalysis
is
the
lack
of
a
"
gold
standard"
with
which
to
compare
the
estimated
TTHM
ingestion
of
subjects
in
the
study.
In
the
absence
of
such
a
standard,
it
is
not
possible
to
determine
whether
the
reanalysis
actually
reduced
exposure
misclassification.
The
conclusions
reached
by
the
study
authors
were
1)
there
was
no
advantage
in
using
the
closest­
site
method
over
the
utility­
wide
method
for
exposure
analysis;
2)
use
of
variance­
based
weighting
factors
and
subset
analyses
is
defensible
and
resulted
in
some
increases
of
odds
ratio,
but
resulting
loss
of
sample
size
may
limit
the
utility
of
these
techniques;
and
3)
use
of
a
variety
of
exposure
assessment
techniques
may
lessen
the
impact
of
Draft
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February
20,
2003
VI
­
10
bias
resulting
from
utility­
specific
factors
such
as
inconsistencies
in
sampling
density
or
unrecognized
contamination
problems.

The
reanalysis
conducted
by
Waller
et
al.
(
2001)
identified
evidence
for
differential
misclassification
in
the
prior
analysis
of
a
ground
water
predominate
area
("
Zone
A")
reported
in
Waller
et
al.
(
1998).
The
effect
of
this
misclassification
was
to
bias
the
original
estimate
of
the
relationship
between
TTHM
ingestion
and
spontaneous
abortion
away
from
the
null.
Over
400
of
the
women
in
the
study
cohort
resided
in
Zone
A,
an
area
within
a
large
mixed­
source
utility
that
received
predominately
groundwater.
Zone
A
was
not
sampled
for
THMs
during
the
study
period.
Because
other
areas
within
the
utility
frequently
had
high
TTHM
concentrations,
use
of
a
utility­
wide
approach
for
estimating
TTHM
concentration
probably
resulted
in
an
overestimation
of
exposure
for
Zone
A
residents.
An
investigation
by
the
study
authors
revealed
that
although
the
spontaneous
abortion
rate
of
women
in
Zone
A
was
low
overall,
women
who
drank
at
least
5
glasses
of
water
per
day
had
a
spontaneous
abortion
rate
of
14.6%.
The
reason
for
the
high
spontaneous
abortion
rate
among
women
consuming
large
amounts
of
Zone
A
water
was
unclear,
but
was
reported
to
be
consistent
with
other
epidemiological
studies
that
found
high
rates
of
spontaneous
abortion
among
women
ingesting
large
amounts
of
unchlorinated
water
in
Region
1
of
the
original
study.
Exclusion
of
Zone
A
residents
or
recoding
them
to
a
level
determined
by
later
testing
within
the
zone
decreased
the
adjusted
OR
for
high
exposure
to
TTHM
(
TTHM

75
µ
g/
L
and
intake

5
glasses
per
day)
to
1.5
(
95%
C.
I.
0.8,
2.8)
as
compared
to
the
adjusted
OR
of
1.8
(
95%
C.
I.
1.1,
3.0)
identified
in
the
original
analysis
(
Waller
et
al.,
1998).
The
impact
of
this
finding
on
the
adjusted
OR
calculated
for
individual
brominated
THMs
is
currently
unknown,
but
is
expected
to
be
addressed
in
a
future
publication
by
Waller
et
al.

King
et
al.
(
2000)
conducted
a
population­
based
retrospective
cohort
study
to
examine
the
relationship
between
TTHM
or
individual
THMs
and
risk
for
stillbirth
of
fetuses
greater
than
500
grams.
The
study
cohort
was
assembled
from
a
perinatal
database
in
Nova
Scotia,
Canada
and
consisted
of
49,756
singleton
births
that
occurred
between
1988
and
1995.
Exposure
was
assigned
by
relating
the
mother's
residence
at
the
time
of
delivery
to
the
levels
of
total
and
individual
THMs
measured
in
public
water
supplies.
Maternal
age,
parity,
smoking
during
pregnancy,
infant's
sex,
and
neighborhood
family
income
were
evaluated
as
potential
cofounders.
Relative
risks
were
adjusted
for
smoking
and
maternal
age.
Exposure
to
TTHMs,
chloroform,
and
bromodichloromethane
were
associated
with
increased
risk
of
stillbirth.
Analysis
of
the
results
suggested
that
exposure
to
bromodichloromethane
was
a
stronger
predictor
of
risk
than
exposure
to
chloroform.
Risk
doubled
for
women
exposed
to
a
bromodichloromethane
level
of
greater
than
or
equal
to
20
µ
g/
L
when
compared
to
women
consuming
concentrations
of
less
than
5
µ
g/
L
(
relative
risk
=
1.98,
95%
C.
I.
1.23,
3.49).
When
categories
of
stillbirth
(
unexplained
deaths
and
asphyxia­
related
deaths)
were
examined,
relative
risk
estimates
for
asphyxia­
related
deaths
increased
by
32%
for
each
10
µ
g/
L
increase
in
concentration
of
bromodichloromethane.
As
noted
by
the
authors,
a
potential
limitation
of
this
study
was
misclassification
of
exposure
as
a
result
of
mobility
within
the
study
population
(
estimated
to
affect
less
than
10%
of
study
subjects).
This
study
did
not
examine
early
fetal
death
(
e.
g.
spontaneous
abortion)
because
the
perinatal
database
employed
in
this
investigation
contained
information
only
on
fetuses
that
weighed
500
grams
or
more.
Draft
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February
20,
2003
VI
­
11
Dodds
and
King
(
2001)
conducted
a
retrospective
cohort
study
of
singleton
births
among
49,842
residents
of
Nova
Scotia,
Canada
between
1988
and
1995
to
assess
the
relationship
between
exposure
to
THMs
and
birth
defects.
Of
the
brominated
THMs,
only
bromodichloromethane
was
examined.
Information
on
exposure
concentrations
consisted
of
routine
water
monitoring
data
obtained
from
the
Nova
Scotia
Department
of
the
Environment
and
reflected
samples
collected
from
within
the
water
distribution
system.
The
birth
defects
examined
had
previously
been
reported
in
other
epidemiological
studies,
and
included
neural
tube
defects,
cardiovascular
defects,
cleft
defects,
and
chromosomal
abnormalities.
The
perinatal
information
used
in
the
study
was
obtained
from
the
Nova
Scotia
Atlee
perinatal
database.
This
database
contains
information
abstracted
from
medical
records
and
includes
infant
diagnoses
among
stillborn
and
live
born
infants
up
to
the
time
of
discharge
from
the
hospital
after
birth.
In
addition,
information
on
prenatally
diagnosed
congential
anomalies
was
obtained
from
pregnancy
terminations.
Inclusion
of
this
data
was
deemed
important
because,
in
Nova
Scotia,
approximately
80%
of
neural
tube
defects
are
detected
antenatally
and
the
pregnancy
is
electively
terminated.
Exposure
windows
were
selected
to
target
the
period
before
or
during
gestation
when
exposure
to
a
potential
developmental
toxicant
or
mutagen
might
have
the
most
profound
effect
on
a
particular
developmental
or
genotoxic
endpoint.
The
selected
windows
were
as
follows:
average
bromoform
concentrations
from
one
month
and
one
month
after
were
used
for
analysis
of
neural
tube
defects;
concentrations
during
the
first
two
months
of
pregnancy
were
used
for
analysis
of
cardiac
defects
and
cleft
defects;
and
the
average
concentrations
three
months
before
pregnancy
were
used
for
the
analysis
of
chromosomal
abnormalities.
Estimates
of
relative
risks
and
95%
confidence
intervals
were
obtained
from
Poisson
regression
models.
Maternal
age,
parity,
maternal
smoking,
and
neighborhood
family
income
were
assessed
as
potential
confounders.
The
categories
used
for
bromodichloromethane
concentration
were
<
5
µ
g/
L;
5
­
9
µ
g/
L;
10
­
19
µ
g/
L;
and

20
µ
g/
L.

Exposure
to
bromodichloromethane
at
concentrations

20
µ
g/
L
was
associated
with
increased
risk
of
neural
tube
defects
(
adjusted
relative
risk
2.5;
95%
C.
I.
1.2,
5.1)
when
adjusted
for
maternal
age
and
income
level.
However,
there
was
no
evidence
of
a
dose­
response
trend
with
increasing
concentration
of
bromodichloromethane.
In
addition,
the
study
authors
noted
that
this
point
estimate
was
"
fairly
unstable"
as
a
result
of
the
low
number
of
cases
(
n=
10)
in
the

20
µ
g/
L
exposure
category.
For
cardiac
defects,
a
significant
reduction
in
risk
was
associated
with
exposure
to
concentrations
of

20
µ
g/
L
(
relative
risk
0.3;
95%
C.
I.
0.2,
0.7)
and
there
was
a
trend
of
decreasing
risk
with
increasing
exposure.
The
study
authors
considered
it
unlikely
that
exposure
above

20
µ
g/
L
was
actually
protective.
They
suggested
that
this
may
be
a
chance
finding
or
a
reflection
of
a
negative
association
of
bromodichloromethane
with
other
disinfection
by­
products
in
this
region
which
may
increase
cardiac
risks.
There
was
no
apparent
trend
or
significant
association
for
exposure
to
bromodichloromethane
and
occurrence
of
cleft
defects
or
chromosomal
aberrations.

2.
Dibromochloromethane
Windham
et
al.
(
2003)
examined
menstrual
cycle
characteristics
in
relation
to
the
presence
of
brominated
trihalomethanes
in
tap
water
in
a
prospective
study
of
women
living
in
Northern
California.
Data
were
also
reported
for
chloroform
and
TTHM.
The
relationships
examined
Draft
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February
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2003
VI
­
12
included:
1)
cycle
characteristics
and
concentration
of
individual
THMs,
TTHM,
and
total
brominated
THMs
in
tap
water;
2)
cycle
characteristics
and
estimated
water
consumption
(
TTHM);
and
3)
cycle
characteristics
and
duration
of
showering
(
TTHM).
The
target
population
was
married
women
of
reproductive
age
(
18­
39
years
old)
who
were
members
of
the
Kaiser
Permanente
Medical
Care
Program.
Participants
in
the
study
were
enlisted
between
May
1990
and
June
1991.
Participants
were
selected
from
among
nearly
6500
women
using
a
short
screening
interview
to
identify
women
who
were
more
likely
to
become
pregnant
(
i.
e.,
those
who
reported
a
menstrual
period
within
the
last
six
weeks,
were
not
surgically
sterilized,
did
not
use
birth
control
pills
or
IUDs,
and
were
non­
contracepting
for
less
than
3
months).
Out
of
1092
eligible
women,
a
total
of
403
women
collected
first
morning
urine
samples
daily
for
2­
9
menstrual
cycles
(
average
5.6
cycles)
for
measurement
of
steroid
metabolites.
These
measurements
were
used
to
estimate
menstrual
parameters
such
as
cycle
and
phase
length.
Cycle
length
was
calculated
from
the
first
day
of
menses
to
the
day
before
onset
of
the
next
menses.
When
the
available
data
permitted,
the
cycle
was
divided
into
the
follicular
phase
(
first
day
of
menses
through
estimated
day
of
ovulation)
and
the
subsequent
luteal
phase.
Between
1424
and
1714
cycles
were
available
for
evaluation
of
each
parameter.
Information
on
water
consumption
(
as
unheated
tap
water
or
drinks
made
from
unheated
tap
water,
drinks
made
from
heated
tap
water,
and
bottled
water)
and
other
variables
(
age,
race,
education,
employment,
income,
pregnancy
history,
exercise
type
and
frequency,
smoking,
alcohol
and
caffeine
consumption)
was
collected
in
a
baseline
telephone
interview
prior
to
urine
collection.
Information
on
the
number
of
showers
taken
at
home
per
week
and
their
duration
was
also
collected.
Showering
was
examined
as
minutes
per
week
and
by
combining
the
duration
and
cycle­
specific
THM
level
to
create
combinations
of
high
and
low
exposure.
The
participants
filled
out
a
daily
diary
during
the
urine
collection
phase
and
recorded
vaginal
bleeding
as
number
of
pads
or
tampons.
Exposure
to
THMs
was
estimated
from
quarterly
utility
monitoring
data
and
information
on
drinking
water
and
other
tap
water
use
collected
during
the
baseline
interview.
A
90­
day
exposure
time
period
was
assigned
for
each
cycle
because
THM
monitoring
was
conducted
by
the
utilities
on
a
quarterly
(
i.
e.,
about
90
days)
basis.
A
period
of
60
days
before
and
30
days
after
each
cycle
start
date
was
selected
for
the
90­
day
window.
Cycle­
specific
exposures
to
TTHM
and
individual
THMs
were
calculated
by
averaging
all
THM
measurements
taken
by
a
participant's
utility
at
various
points
in
the
distribution
system
(
i.
e.,
the
"
utility­
wide
average"
described
by
Waller
et
al.,
1998,
2001)
during
the
that
90­
day
period.
Because
the
brominated
THMs
were
highly
correlated
and
thus
difficult
to
examine
independently,
the
study
authors
also
examined
the
sum
of
the
levels
of
the
three
brominated
compounds.
Exposure
levels
for
the
brominated
THMs
were
examined
as
categorical
variables.
For
TTHM,
ingestion
metrics
were
calculated
for
unheated
tap
water
and
the
sum
of
heated
and
unheated
tap
water.
Statistical
analyses
were
conducted
using
the
menstrual
cycle
as
the
unit
of
observation.
Menstrual
parameters
were
analyzed
as
continuous
or
categorical
variables
in
relation
to
categorical
exposure
indices
and
the
methods
used
accounted
for
repeated
measures.
Numerous
covariates
reflecting
demographic,
reproductive
history,
and
lifestyle
factors
were
examined
in
relation
to
categorical
trihalomethane
levels
and
ingestion
to
identify
potential
confounders.
Age,
pregnancy
history,
body
mass
index,
caffeine
consumption,
and
alcohol
consumption,
as
well
as
race
and
smoking,
were
included
as
variables
in
all
adjusted
models.
Draft
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February
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2003
VI
­
13
Increasing
levels
of
individual
brominated
compounds
or
total
brominated
THMs
were
associated
with
significantly
shorter
cycles
when
examined
by
quartile
(
Table
VII­
3).
Similar
decrements
were
observed
in
follicular,
but
not
luteal,
phase
length.
Dose­
response
patterns
were
evident
for
both
individual
and
total
brominated
THMs.
The
strongest
association
for
an
individual
compound
was
observed
for
dibromochloromethane,
with
adjusted
decrements
of
1.2
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2003
VI
­
14
Table
VI­
3
Means
and
Adjusted
Differences
in
Menstrual
Cycle
and
Follicular
Phase
Length
by
Quartile
of
Individual
and
Summed
Brominated
Trihalomethanes
Compound
Quartile
of
Exposurea
1b
2­
3
4
Mean
in
days
(
SE)
Adjusted
Differencec
(
95%
CI)
Adjusted
Difference
(
95%
CI)

Cycle
Length
Bromodichloromethane
29.8
(
0.30)
­
0.59
(­
1.2,
­
0.02)
­
0.74
(­
1.5,
­
0.02)

Dibromochloromethane
30.0
(
0.33)
­
0.69
(­
1.4,
­
0.02)
­
1.2
(­
2.0,
­
0.38)

Bromoform
29.7
(
0.26)
­
0.42
(­
0.96,
0.13)
­
0.79
(­
1.4,
­
0.14)

Sum
of
Brominated
Compounds
30.0
(
0.34)
­
0.72
(­
1.4,
­
0.04)
­
1.2
(­
2.0,
­
0.40)

Follicular
Phase
Bromodichloromethane
17.0
(
0.31)
­
0.54
(­
1.1,
0.01)
­
0.80
(­
1.5,
­
0.08)

Dibromochloromethane
17.1
(
0.34)
­
0.62
(­
1.3,
0.05)
­
1.1
(­
1.9,
­
0.25)

Bromoform
16.9
(
0.27)
­
0.30
(­
0.83,
0.23)
­
0.78
(­
1.4,
­
0.14)

Sum
of
Brominated
Compounds
17.2
(
0.35)
­
0.66
(­
1.3,
0.02)
­
1.1
(­
1.9,
­
0.29)

Source:
Windham
et
al.
(
2003)
a
Top
quartiles
for
bromodichloromethane,
dibromochloromethane,
bromoform,
and
the
summed
brominated
compounds
are

16,

20,

12,
and

45
µ
g/
L,
respectively.
b
Reference
group;
the
mean
provided
is
unadjusted
with
standard
error
(
SE)
c
Adjusted
for
age,
race,
body
mass
index,
income,
pregnancy
history,
caffeine
and
alcohol
consumption,
and
smoking.

days
(
95%
C.
I.
­
2.0,
­
0.38)
for
mean
cycle
length
and
1.1
days
(
95%
C.
I.
­
1.9,
­
0.25)
for
mean
follicular
phase
length
at
the
highest
quartile
(

20
µ
g/
L).
In
comparison,
a
clear
association
with
reduced
cycle
length
was
not
observed
for
chloroform
(
difference
­
0.3
days;
95%
C.
I.
­
1.0,
0.40),
even
at
the
highest
quartile
(

17
µ
g/
L).
Menses
length
was
slightly
increased
at
the
highest
quartile
for
bromodichloromethane
exposure.
For
categorical
variables,
the
odds
of
having
a
long
cycle
(
Adjusted
Odds
Ratio,
AOR
0.55;
95%
C.
I.
0.28,
1.08)
or
long
follicular
phase
(
AOR
0.26;
95%
C.
I.
0.12,
0.59)
were
significantly
reduced
at
the
highest
quartile
for
summed
brominated
THM
concentration
(

45
µ
g/
L).
These
data
suggest
that
brominated
THMs
or
other
disinfection
by­
products
that
co­
occur
with
brominated
THMs
may
affect
ovarian
function.
Draft
­
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Not
Cite
or
Quote
February
20,
2003
VI
­
15
Windham
et
al.
(
2003)
also
examined
relationships
between
TTHM
(
brominated
compounds
plus
chloroform)
exposure
and
menstrual
parameters.
A
monotonic
decrease
in
mean
cycle
length
was
observed
with
increasing
TTHM
exposure
category.
At
concentrations
greater
than
60
µ
g/
L,
the
adjusted
decrement
was
1.1
day
(
95%
C.
I.
­
1.8,
­
0.40)
when
compared
to
concentrations
less
than
40
µ
g/
L.
The
decrease
in
follicular
phase
length
was
similar
(­
0.94
day;
95%
C.
I.
­
1.6,
­
0.24).
Cycles
with
TTHM
concentrations
above
the
current
MCL
of
80
µ
g/
L
were
appeared
to
be
shorter
by
about
one
day
(
0.99
days;
95%
C.
I.
­
2.2,
0.17).
A
unit
decrement
in
mean
cycle
and
follicular
phase
length
of
0.18
days
per
10
µ
g/
L
increase
in
total
trihalomethane
concentration
(
95%
C.
I.
­
0.29,
­
0.07)
was
determined
when
the
cycle­
specific
TTHM
level
was
examined
as
a
continuous
variable.
When
ingestion
patterns
were
examined,
mean
cycle
and
phase
lengths
showed
little
variation
in
relation
to
consumption
of
unheated
tap
water
at
home.
In
contrast,
increased
consumption
of
heated
tap
water
was
associated
with
significantly
decreased
cycle
and
follicular
phase
lengths.
The
observed
decrements
were
greater
than
one
day
with
daily
consumption
of
three
or
more
drinks
made
from
heated
tap
water.
These
decrements
were
reduced
by
adjustment,
particularly
when
caffeine
was
included
in
the
model;
the
adjusted
decrement
in
cycle
length
was
0.68
days
(
95%
C.
I.
­
2.1,
0.72).
A
non­
monotonic
relationship
was
observed
for
mean
cycle
length
and
an
ingestion
metric
combining
TTHM
concentration
and
consumption
of
unheated
tap
water,
with
the
highest
category
(>
60
µ
g/
L)
showing
a
decrement
of
0.4
days
and
the
third
category
(>
40­
60
µ
g/
L)
showing
a
decrement
of
one
day.
Use
of
an
ingestion
metric
based
on
total
home
tap
water
consumption
(
i.
e.,
heated
and
unheated
tap
water)
revealed
a
more
consistent
pattern
of
reduced
cycle
length,
with
adjusted
decrements
of
greater
than
one
day
for
cycle
(­
1.1
day;
95%
C.
I.
­
2.2,
­
0.06)
and
follicular
phase
(­
1.1
day;
95%
C.
I.
­
2.2,
0.03)
lengths.
Examination
of
time
spent
showering
did
not
reveal
additional
risks
with
longer
showers.
The
unadjusted
mean
cycle
length
varied
little
by
time
spent
showering.
Following
adjustment,
there
was
a
tendency
toward
decreased
length
with
any
category
of
showering
above
35
minutes/
week.
This
relationship
was
stronger
for
follicular
phase
duration
than
for
cycle
length.
For
example,
the
adjusted
mean
decrements
at
the
longest
duration
(

105
minutes)
were
­
0.68
days
(
95%
C.
I.
­
2.0,
0.68)
for
cycle
length
and
­
1.2
days
(
95%
C.
I.
­
2.6,
0.26).
However,
the
confidence
intervals
were
wide
for
all
duration
categories
and
a
clear
dose
response
pattern
(
i.
e.,
shorter
lengths
at
higher
durations)
was
not
evident.

The
study
authors
noted
several
strengths
and
potential
limitations
of
this
study.
Strengths
include
the
use
of
a
prospective
study
design,
use
of
biologic
measures
to
determine
menstrual
parameters,
and
consideration
of
many
potential
confounders.
Potential
limitations
include
exposure
misclassification,
use
of
a
study
sample
that
is
not
representative
of
the
general
population,
lack
of
information
on
other
sources
of
exposure
such
as
washing,
cooking
and
cleaning,
as
well
as
exposures
outside
the
home.
There
are
two
observations
in
this
study
that
might
suggest
involvement
of
compounds
other
than
THMs
in
the
reduction
of
cycle
length.
First,
the
more
consistent
association
of
decreased
cycle
length
reported
for
heated
compared
to
unheated
tap
water
is
unexpected
if
THMs
alone
are
the
causative
agent.
This
is
because
THMs
volatilize
from
heated
water
and
exposure
to
these
compounds
should
therefore
be
lower
for
heated
tap
water,
unless
the
volatilized
compound
is
inhaled.
Second,
examination
of
time
spent
showering
did
not
reveal
additional
risks
with
longer
showers.
This
is
also
counter
to
the
expected
trend,
as
elevated
blood
levels
of
THMs
have
been
documented
after
showering
(
Backer
et
al.,
2000;
Lynberg
et
al.,
2001)
as
a
result
of
dermal
and
inhalation
exposure.
However,
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VI
­
16
information
on
shower
duration
was
collected
by
interview
and
the
reported
lengths
may
not
have
accurately
reflected
actual
shower
duration.

Because
the
study
by
Windham
et
al.
(
2003)
is
the
first
to
examine
changes
in
menstrual
cycle
function
in
relation
to
tap
water
exposure,
there
are
no
supporting
data
on
the
association
of
disinfection
by­
products
other
than
the
THMs
with
changes
in
menstrual
cycle
function.
Although
this
study
suggests
that
disinfection
by­
products
may
have
effects
on
ovarian
function,
no
conclusions
can
be
drawn
regarding
the
identity
of
the
compounds
responsible
for
these
effects.
The
relationships
reported
in
this
study
underscore
the
need
for
additional
research
on
the
effects
of
drinking
water
disinfection
by­
products
on
menstrual
cycle
function.

3.
Bromoform
Epidemiological
studies
identifying
adverse
health
effects
specifically
associated
with
exposure
to
bromoform
were
not
identified.

C.
High
Risk
Populations
High
risk
(
or
susceptible)
populations
are
those
which
experience
more
adverse
effects
at
comparable
levels
of
exposure,
which
experience
adverse
effects
at
lower
exposure
levels
than
the
general
population,
or
which
experience
a
higher
than
average
exposure
because
they
live
or
work
in
settings
with
elevated
environmental
concentrations
of
the
chemical
of
interest.
The
enhanced
response
of
these
susceptible
subpopulations
may
result
from
intrinsic
or
extrinsic
factors.
Factors
that
may
be
important
include,
but
are
not
limited
to:
impaired
function
of
detoxification,
excretory,
or
compensatory
processes
that
protect
against
or
reduce
toxicity;
differences
in
physiological
protective
mechanisms;
genetic
differences
in
metabolism;
developmental
stage;
health
status;
gender;
or
age
of
the
individual.
For
brominated
trihalomethanes,
high
risk
populations
may
potentially
include
those
with
elevated
levels
of
CYP2E1
(
via
exposure
to
inducing
substances
or
because
of
altered
physiological
or
health
states)
or
elevated
levels
of
glutathione­
S­
transferase
theta.
These
factors
are
discussed
in
greater
detail
in
Section
VII.
D.
3
of
this
document.

A
growing
body
of
scientific
evidence
indicates
that
children
may
suffer
disproportionately
from
some
environmental
health
risks.
These
risks
may
arise
because
the
neurological,
immunological,
and
digestive
systems
of
children
are
still
developing
(
U.
S.
EPA,
(
1998a).
In
addition,
children
may
incur
greater
exposure
because
they
eat
more
food,
consume
more
fluids,
and
breathe
more
air
in
proportion
to
their
body
weight
when
compared
to
adults
(
U.
S.
EPA,
1998a).
Factors
contributing
to
potentially
greater
risk
in
children
are
discussed
in
Section
VII.
D.
2
of
this
document.
Draft
­
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or
Quote
February
20,
2003
VI
­
17
D.
Summary
Limited
human
health
data
are
available
for
the
brominated
trihalomethanes.
In
the
past,
bromoform
was
used
as
a
sedative
for
children
with
whooping
cough.
Doses
of
50
to
100
mg/
kgday
usually
produced
sedation
without
apparent
adverse
effects.
Some
rare
instances
of
death
or
near­
death
were
reported,
although
these
cases
were
generally
attributed
to
accidental
overdoses.
No
human
toxicological
data
were
available
for
bromodichloromethane
or
dibromochloromethane.

Numerous
epidemiological
studies
have
examined
the
association
between
water
chlorination
and
increased
cancer
mortality
rates.
None
of
these
studies
has
examined
the
association
between
cancer
and
exposure
to
any
individual
brominated
trihalomethane.
Recent
studies
have
examined
the
association
of
chlorinated
water
use
with
various
pregnancy
outcomes,
including
low
birth
weight,
premature
birth,
intrauterine
growth
retardation,
spontaneous
abortion,
stillbirth
,
and
birth
defects.
An
association
has
been
reported
for
exposure
to
bromodichloromethane
(
or
a
closely
associated
compound)
and
a
moderately
increased
risk
of
spontaneous
abortion
during
the
first
trimester.
A
confirmation
of
this
finding
is
pending
reanalysis
of
the
original
data
to
correct
a
differential
misclassification
error
identified
in
a
subsequent
analysis
of
the
study
data.
An
association
has
also
been
reported
for
exposure
to
bromodichloromethane
(
or
a
closely
associated
compound)
1)
and
stillbirth
of
fetuses
weighing
more
than
500
g
and
2)
increased
risk
of
neural
tube
defects
in
women
exposed
to

20
µ
g/
L
of
bromodichloromethane
prior
to
conception
through
the
first
month
of
pregnancy.
An
association
has
been
reported
for
total
brominated
trihalomethanes
and
reduced
menstrual
cycle
and
follicular
phase
length
in
women
of
child­
bearing
age.
Among
the
individual
brominated
trihalomethanes,
dibromochloromethane
displayed
the
strongest
association
with
altered
menstrual
function.
It
is
not
possible
to
directly
conclude
from
these
studies
that
bromodichloromethane
and
dibromochloromethane
are
developmental
toxicants
in
humans,
because
chlorinated
water
contains
many
disinfection
by­
products.
Nevertheless,
these
studies
raise
significant
concern
for
possible
human
health
effects.
The
methodology
used
to
estimate
exposure
to
brominated
trihalomethanes
in
tap
water
has
been
examined
with
the
goal
of
refining
estimates
of
intake
of
these
compounds
in
epidemiological
studies.

For
the
brominated
trihalomethanes,
populations
at
high
risk
may
potentially
include
those
with
elevated
levels
of
CYP2E1
(
via
exposure
to
inducing
substances
or
because
of
altered
physiological
or
health
states)
or
elevated
levels
of
glutathione­
S­
transferase
theta.
In
addition,
users
of
hot
tubs
and
swimming
pools
may
experience
additional
exposure
to
brominated
trihalomethanes.
Draft
­
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or
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February
20,
2003
VII
­
1
VII.
MECHANISM
OF
TOXICITY
A.
Role
of
Metabolism
The
toxicity
of
the
brominated
trihalomethanes
is
related
to
their
metabolism.
This
conclusion
is
based
largely
on
the
observation
that
liver
and
kidney,
the
chief
target
tissues
for
these
compounds,
are
also
the
primary
sites
of
their
metabolism.
In
addition,
treatments
which
increase
or
decrease
metabolism
also
tend
to
increase
or
decrease
trihalomethane­
induced
toxicity
in
parallel.
Pankow
et
al.
(
1997),
for
example,
examined
the
relationship
between
metabolism
of
dibromochloromethane
and
hepatotoxicity.
Serum
leucine
aminopeptidase
(
LAP)
activity
(
an
indicator
of
hepatotoxicity)
increased
in
a
dose­
dependant
fashion
with
any
treatment
that
increased
the
metabolism
of
dibromochloromethane
(
e.
g.
pretreatment
with
isoniazid
or
mxylene

B.
Biochemical
Basis
of
Toxicity
The
precise
biochemical
mechanisms
which
link
brominated
trihalomethane
metabolism
to
toxicity
are
not
certain,
but
many
researchers
have
proposed
that
toxicity
results
from
the
production
of
reactive
intermediates.
These
reactive
intermediates
are
believed
to
form
covalent
adducts
with
various
cellular
molecules
and
to
impair
the
function
of
those
molecules
resulting
cell
injury.
Reactive
intermediates
may
arise
from
the
oxidative
(
dihalocarbonyls)
or
the
reductive
(
free
radicals)
pathways
of
metabolism
as
discussed
in
Section
III.
C.
Support
for
this
mode
of
action
has
been
obtained
from
in
vitro
studies
of
bromodichloromethane.
Under
both
aerobic
and
anoxic
conditions,
bromodichloromethane
is
metabolized
to
intermediates
that
covalently
bind
to
rat
microsome
proteins
and
lipids.
Direct
evidence
showing
a
relationship
between
the
levels
of
covalent
binding
intermediates
generated
by
the
oxidative
or
reductive
pathways
and
the
extent
of
toxicity
is
not
currently
available
for
brominated
trihalomethanes.

Free
radical
generation
by
the
reductive
pathway
for
brominated
trihalomethane
metabolism
may
result
in
lipid
peroxidation.
Although
evidence
demonstrating
that
lipid
peroxidation
actually
accounts
for
the
observed
cellular
toxicity
associated
with
brominated
trihalomethanes
is
lacking,
at
least
one
study
has
established
that
lipid
peroxidation
does
occur
in
conjunction
with
brominated
trihalomethane
metabolism.
De
Groot
and
Noll
(
1989)
reported
that
all
three
brominated
trihalomethanes
induced
lipid
peroxidation
in
rat
liver
microsomes
in
vitro,
and
that
this
was
maximal
at
low
oxygen
levels
(
between
1
and
10
mm
Hg
of
O
2).
The
authors
interpreted
these
data
to
support
the
concept
that
lipid
peroxidation
is
caused
by
free
radical
metabolites
generated
by
the
reductive
metabolism
of
trihalomethanes.

Glutathione
has
been
implicated
in
both
defense
against
toxicity
induced
by
brominated
trihalomethanes
and
in
generation
of
mutagenic
metabolites.
Gao
et
al.
(
1996)
examined
the
effect
of
glutathione
on
the
toxicity
of
bromodichloromethane
in
vivo
and
in
vitro.
Depletion
of
glutathione
by
pretreatment
of
male
F344
rats
with
the
glutathione
synthesis
inhibitor
buthionine
sulfoximine
increased
the
hepatotoxicity
of
a
single
gavage
dose
of
400
mg/
kg
bromodichloromethane
administered
in
10%
Emulphor
®
.
Biochemical
indicators
of
hepatotoxicity
(
e.
g.
AST,
ALT,
LDH)
were
increased
approximately
11­
fold
and
the
severity
of
Draft
­
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or
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February
20,
2003
VII
­
2
morphological
changes
(
centrilobular
vacuolar
degeneration
and
hepatocellular
necrosis)
was
greater
in
the
glutathione­
depleted
animals.
Serum
and
urinary
markers
of
renal
damage
were
also
significantly
increased
by
glutathione
depletion.
Renal
tubule
necrosis
was
observed
only
in
the
glutathione­
depleted
group.
Overall,
glutathione
depletion
appeared
to
enhance
hepatotoxicity
more
than
nephrotoxicity,
an
effect
that
was
attributed
to
organ­
specific
differences
in
bromodichloromethane
metabolism.
The
addition
of
glutathione
to
reaction
mixtures
of
rat
hepatic
or
renal
microsomal
fraction
and
radiolabeled
bromodichloromethane
resulted
in
92%
and
20%
reductions
in
protein
binding
to
bromodichloromethane,
respectively.
The
difference
in
response
to
glutathione
addition
was
interpreted
as
evidence
for
existence
of
different
metabolic
pathways
in
liver
and
kidney.
Bromodichloromethane
binding
to
lipid
in
liver
microsomes
under
anaerobic
conditions
was
decreased
in
the
presence
of
glutathione,
suggesting
that
glutathione
can
react
with
the
dihalomethyl
radical.

In
contrast
to
the
apparent
role
protective
role
of
glutathione
described
above,
studies
in
strains
of
S.
typhimurium
engineered
to
express
rat
theta
class
glutathione
S­
transferase
suggest
that
conjugation
with
glutathione
leads
to
formation
of
mutagenic
metabolites
(
Pegram
et
al.,
1997;
DeMarini
et
al.,
1997).
These
studies
are
described
in
greater
detail
in
Section
V.
F.
1.
Proposed
pathways
for
generation
of
the
mutagenic
species
are
outlined
in
Figure
V­
2,
also
located
in
Section
V.
F.
1.
Briefly,
similar
mutational
specificity,
site
specificity,
and
mutation
spectra
for
the
three
brominated
trihalomethanes
support
the
conclusion
that
they
are
activated
by
one
or
more
common
pathways.
In
contrast,
the
data
do
not
support
a
glutathione
S­
transferase
mediated
pathway
for
the
structurally­
related
trihalomethane
chloroform.
This
finding
suggests
that
chloroform
and
the
brominated
trihalomethanes
may
in
some
instances
be
metabolized
by
different
pathways.

C.
Mode
of
Action
of
Carcinogenesis
Administration
of
individual
brominated
trihalomethanes
has
been
associated
with
formation
of
liver
tumors
(
bromodichloromethane,
dibromochloromethane),
kidney
tumors
(
bromodichloromethane),
and
tumors
of
the
large
intestine
(
bromodichloromethane,
bromoform)
in
some
experimental
animals.
The
mode
of
action
by
which
brominated
trihalomethanes
induce
tumors
in
laboratory
animals
is
not
known.
However,
two
general
modes
of
action
have
been
proposed:
1)
formation
of
DNA
adducts
resulting
from
interaction
with
one
or
more
classes
of
reactive
metabolites
and
2)
production
of
cytotoxicity
coupled
with
regenerative
hyperplasia.

The
production
of
reactive
metabolites
from
trihalomethanes
is
well­
established.
Classes
of
reactive
metabolites
produced
include
dihalocarbonyls
produced
by
oxidative
metabolism
and
and
dihalomethyl
radicals
produced
by
reductive
metabolism.
Additional
evidence
suggests
that
reactive
species
can
be
also
formed
via
glutathione
conjugation
(
DeMarini
et
al.,
1997;
Pegram
et
al.,
1997).
Detection
of
adduct
formation
and
consistent
evidence
of
DNA
reactivity
in
standard
assays
are
two
lines
of
experimental
evidence
that
would
strongly
support
the
adduct
formation
hypothesis.
At
present
there
are
no
in
vivo
data
available
on
DNA
adducts
resulting
from
metabolism
of
brominated
trihalomethanes.
DNA
reactivity
can
be
inferred
from
test
results
of
mutagenic
and
genotoxic
potential.
As
noted
previously
(
U.
S.
EPA,
1994b),
synthesis
of
the
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VII
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overall
weight
of
evidence
for
genotoxicity
of
individual
brominated
trihalomethanes
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid),
and
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
Overall,
a
majority
of
studies
yielded
more
positive
results
for
bromoform
and
bromodichloromethane,
and
evidence
of
mutagenicity
is
considered
adequate
for
these
chemicals.
Study
results
for
the
mutagenicity
of
dibromochloromethane
are
mixed,
and
the
overall
evidence
for
mutagenicity
of
this
chemical
is
judged
to
be
inconclusive
(
U.
S.
EPA,
1994b).

Alternatively,
the
induction
of
tumors
by
individual
brominated
trihalomethanes
could
involve
an
epigenetic
mode
of
action.
Induction
of
tumors
in
animal
studies
has
been
noted
to
occur
primarily
at
sites
where
cytotoxicity
was
observed
(
i.
e.,
liver
and
kidney),
and
there
appears
to
be
a
correlation
between
hepatotoxicity
and
liver
tumorigenicity
of
brominated
trihalomethanes
in
mice
(
bromodichloromethane
>
dibromochloromethane
>
bromoform)
(
U.
S.
EPA,
1994b).
This
raises
the
possibility
that
regenerative
hyperplasia
caused
by
the
cytotoxic
effects
of
brominated
trihalomethanes
may
contribute
to
the
tumorigenic
potential
of
these
chemicals.
A
brief
review
of
studies
that
have
evaluated
regenerative
hyperplasia
following
exposure
to
brominated
trihalomethanes
is
provided
below.

A
number
of
studies
have
measured
cell
proliferation
in
liver
and/
or
kidney
following
exposure
to
brominated
trihalomethanes.
Miyagawa
et
al.
(
1995)
observed
evidence
for
the
induction
of
hepatocyte
proliferation
in
male
B6C3F
1
mice
following
a
single
oral
gavage
dose
of
dibromochloromethane
in
corn
oil
at
the
maximum
tolerated
dose
(
MTD)
or
at
one
half
the
MTD
(
200
or
400
mg/
kg).
Proliferation
was
assessed
by
incorporation
of
[
3H]­
thymidine
using
the
in
vivo­
in
vitro
replicative
DNA
synthesis
assay
at
24,
39,
and
48
hours
postexposure.

Potter
et
al.
(
1996)
investigated
cell
proliferation
in
the
kidney
of
male
F344
rats.
Test
animals
received
0.75
or
1.5
mmol/
kg
of
bromodichloromethane
in
4%
Emulphor
®
by
gavage
for
1,
3,
or
7
days.
The
administered
doses
corresponded
to
123
or
246
mg/
kg­
day
for
bromodichloromethane,
156
or
312
mg/
kg­
day
for
dibromochloromethane,
and
190
or
379
mg/
kg­
day
for
bromoform.
Cell
proliferation
in
the
kidney
was
assessed
in
vivo
by
[
3H]­
thymidine
incorporation.
No
statistically
significant
effect
of
bromodichloromethane
on
tubular
cell
proliferation
was
observed
following
exposures
of
up
to
7
days,
although
high
labeling
levels
were
observed
in
3
of
4
rats
at
the
246
mg/
kg­
day
dose
of
bromodichloromethane.

NTP
(
1998)
evaluated
cell
proliferation
in
the
kidney
and
liver
of
Sprague­
Dawley
rats
as
part
of
a
short­
term
reproductive
and
developmental
toxicity
screen
of
bromodichloromethane.
The
compound
was
administered
in
drinking
water
for
35
days.
Groups
of
male
and
female
rats
were
exposed
to
drinking
water
concentrations
of
0,
100,
700
and
1300
ppm
bromodichloromethane
using
the
study
design
described
in
Table
V­
6
(
Section
V.
D.
1).
BrdU
labeling
index
(
LI)
was
unchanged
in
the
livers
and
kidneys
of
Group
B
males
at
doses
up
to
69
mg/
kg­
day.
A
small
but
statistically
significant
increase
in
the
LI
was
noted
in
the
livers
and
kidneys
of
Group
C
females
in
the
1300
ppm
dose
group
(
126
mg/
kg­
day).
The
study
authors
noted
that
the
result
in
females
was
probably
biologically
insignificant.
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20,
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VII
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Melnick
et
al.
(
1998)
exposed
female
B6C3F
1
mice
(
10/
dose)
to
bromodichloromethane,
dibromochloromethane,
or
bromoform
in
corn
oil
via
gavage
for
3
weeks
(
5
days/
week).
BrdU
was
administered
to
the
animals
during
the
last
6
days
of
the
study,
and
hepatocyte
labeling
index
(
LI)
analysis
was
conducted.
Time­
adjusted
doses
of
107,
336,
and
357
mg/
kg
of
bromodichloromethane,
dibromochloromethane,
and
bromoform,
respectively,
resulted
in
significantly
elevated
hepatocyte
proliferation
as
measured
by
the
LI.
The
authors
compared
the
dose
response
for
liver
toxicity
(
including
hepatic
enzyme
activity
and
LI
data)
and
tumorigenicity
(
using
data
from
previously
published
NTP
bioassays)
for
the
brominated
trihalomethanes
using
the
Hill
equation
model.
This
analysis
indicated
that
the
shape
of
the
dose
response
as
well
as
the
Hill
exponents
were
different
for
liver
toxicity
and
tumorigenicity.
The
authors
concluded
that
these
data
do
not
support
a
causal
relationship
between
liver
toxicity/
reparative
hyperplasia
and
tumor
development.

Torti
et
al.
(
2001)
conducted
1­
week
and
3­
week
inhalation
exposure
studies
of
bromodichloromethane
in
wild
type
and
transgenic
mice.
Bromodichloromethane
toxicity
was
transient.
Regenerative
lesions
and
increased
labeling
index
were
evident
in
the
kidney
cortex
of
mice
exposed
to
concentrations
of
10
ppm
and
above
for
one
week.
After
three
weeks
of
bromodichloromethane
exposure,
damaged
areas
of
kidney
cortex
were
entirely
regenerated
(
residual
scarring
was
present)
and
labeling
index
measurements
had
returned
to
near
baseline
levels.
The
study
authors
noted
that
these
results
are
in
contrast
to
those
observed
in
similar
experiments
performed
with
chloroform,
where
treatment
of
F344
rats
and
B6C3F
1
mice
resulted
in
continued
cytotoxicity
and
elevated
cell
turnover
for
up
to
90
days
(
Larson
et
al.,
1996;
Templin
et
al.,
1996).
The
mechanistic
basis
for
these
different
responses
to
structurally
similar
compound
is
unclear,
but
may
reflect
an
induced
change
in
metabolism
or
emergence
of
a
resistant
cell
population
in
animals
treated
with
bromodichloromethane.

George
et
al.
(
2002)
reported
that
exposure
of
male
F344/
N
rats
to
bromodichloromethane
in
drinking
water
for
two
years
at
a
level
that
significantly
enhanced
the
prevalence
and
multiplicity
of
hepatocellular
adenomas
and
carcinomas
had
no
effect
on
hepatocellular
proliferation.
In
the
same
study,
the
prevalence
of
renal
tubular
hyperplasia,
but
not
tumor
incidence,
was
significantly
increased
at
the
high
dose.

Based
on
an
extensive
evaluation
of
carcinogenicity
data,
cytotoxicity
coupled
with
regenerative
hyperplasia
is
considered
the
primary
mode
of
action
for
tumor
formation
following
exposure
to
high
concentrations
of
chloroform,
a
structurally­
related
trihalomethane
which
has
low
genotoxic
potential
(
U.
S.
EPA,
2000d).
However,
two
lines
of
evidence
suggest
that
chloroform
is
not
a
prototypical
trihalomethane.
First,
the
weight­
of­
evidence
for
at
least
two
of
the
brominated
trihalomethanes
indicates
that
they
are
genotoxic.
This
contrasts
with
the
negative
weight
of
evidence
evaluation
for
chloroform.
Second,
there
is
evidence
that
the
brominated
trihalomethanes
are
readily
bioactivated
to
mutagenic
products
via
a
glutathione
S­
transferase
mediated
pathway,
while
chloroform
is
bioactivated
only
at
very
high
concentrations.
Therefore,
a
common
mode
of
action
for
carcinogenicity
of
chloroform
and
brominated
trihalomethanes
cannot
be
assumed
on
the
basis
of
current
experimental
evidence.
Data
to
support
a
nonlinear
primary
mode
of
action
for
tumor
development
in
liver,
kidney,
and
large
intestine
are
currently
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VII
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5
lacking
for
the
brominated
trihalomethanes.
In
the
absence
of
such
information,
combined
with
a
positive
weight­
of­
evidence
evaluation
for
genotoxicity,
the
mode
of
action
for
tumor
development
is
assumed
to
be
a
linear
process.

D.
Interactions
and
Susceptibilities
1.
Potential
Interactions
The
toxicity
of
the
brominated
trihalomethanes
appears
to
result
from
cytochrome
P450­
mediated
metabolism
to
reactive
metabolites
(
U.
S.
EPA,
1994b).
Therefore,
agents
which
increase
or
decrease
the
activity
of
enzymes
responsible
for
metabolism
of
brominated
trihalomethanes
may
modify
toxicity.
Pankow
et
al.
(
1997)
observed
that
pretreatment
with
isoniazid
or
m­
xylene
(
inducers
of
CYP2E1
and
CYP2B1/
CYP2B2,
respectively)
increased
the
hepatotoxicity
of
dibromochloromethane
in
male
rats,
as
measured
by
elevated
serum
leucine
aminopeptidase
activity.
Hewitt
et
al.
(
1983)
observed
that
pretreatment
with
acetone,
a
CYP2E1
inducer,
potentiated
the
acute
toxicity
of
bromodichloromethane
and
dibromochloromethane
in
male
rats.
Thornton­
Manning
et
al.
(
1993)
also
found
that
pretreatment
with
acetone
potentiated
the
acute
hepatotoxicity
of
bromodichloromethane
in
male
rats.
Conversely,
the
cytochrome
P450
inhibitor
1­
aminobenzotriazole
prevented
bromodichloromethane­
induced
hepatotoxicity
in
rats
(
Thornton­
Manning
et
al.
1993).
Current
findings
regarding
the
existence
of
glutathionemediated
pathways
for
brominated
trihalomethane
metabolism
(
see
sectionV.
E.
1)
suggest
that
treatments
or
agents
which
alter
glutathione­
S­
transferase
activity
may
potentially
modify
the
toxicity
of
brominated
trihalomethanes.

The
severity
of
brominated
trihalomethane
toxicity
is
potentially
affected
by
the
vehicle
of
administration.
Vehicle
effects
are
well­
documented
in
the
toxicity
of
chloroform
(
e.
g.,
Bull
et
al.
1986;
Jorgenson
et
al.
1985)
and
there
is
some
evidence
that
similar
effects
occur
with
brominated
trihalomethanes.
In
a
study
of
vehicle
effects
on
the
acute
toxicity
of
bromodichloromethane,
a
high
dose
(
400
mg/
kg)
of
the
chemical
was
more
hepato­
and
nephrotoxic
when
given
in
corn
oil
compared
to
aqueous
administration,
but
this
difference
was
not
evident
at
a
lower
dose
(
200
mg/
kg)
(
Lilly
et
al.
1994).
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20,
2003
VII
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6
2.
Greater
Childhood
Susceptibility
A
growing
body
of
scientific
evidence
indicates
that
children
may
suffer
disproportionately
from
some
environmental
health
risks.
These
risks
may
arise
because
the
neurological,
immunological,
and
digestive
systems
of
children
are
still
developing
(
U.
S.
EPA,
(
1998a).
In
addition,
children
may
incur
greater
exposure
because
they
eat
more
food,
consume
more
fluids,
and
breathe
more
air
in
proportion
to
their
body
weight
when
compared
to
adults
(
U.
S.
EPA,
1998a).

U.
S.
EPA
(
1998a)
recently
identified
three
key
questions
to
consider
when
evaluating
health
risks
in
children
from
exposure
to
drinking
water
disinfection
byproducts
(
DBP)
such
as
the
brominated
trihalomethanes:

°
Is
there
information
which
shows
that
the
disinfectant
or
DBP
causes
effects
in
the
developing
fetus
or
impairs
ability
to
conceive
and
bear
children?
If
it
causes
this
type
of
problem
will
it
occur
at
a
lower
dose
than
that
which
will
cause
other
types
of
effects?

°
If
the
disinfectant
or
DBP
causes
cancer,
are
children
more
likely
to
be
affected
by
it
than
are
adults?

°
If
the
disinfectant
or
DBP
causes
some
noncancer
toxic
effect,
are
children
more
likely
to
be
affected
by
it
than
are
adults?

The
data
available
for
evaluation
of
these
issues
as
they
relate
to
brominated
trihalomethanes
are
addressed
below.

a.
Effects
on
the
fetus
and
ability
to
conceive
and
bear
children
General
Results
from
Animal
Studies
Studies
on
the
reproductive
and
developmental
effects
of
brominated
trihalomethanes
are
summarized
in
Section
V.
E.
At
the
present
time,
the
available
data
suggest
that
dibromochloromethane
or
bromoform
cause
effects
in
the
developing
fetus
only
at
doses
which
also
produce
signs
of
maternal
toxicity.
Furthermore,
there
is
no
evidence
that
indicates
that
either
of
these
chemicals
impairs
ability
to
conceive
and
bear
children.
Studies
of
these
chemicals
in
animals
indicate
that
reproductive
and
developmental
effects
would
likely
occur
only
at
doses
higher
than
those
observed
to
cause
liver
and
renal
effects.

Bromodichloromethane
has
the
most
extensive
database
for
developmental
and
reproductive
effect
among
the
brominated
trihalomethanes.
Study
results
for
the
reproductive
and
developmental
effects
of
bromodichloromethane
are
mixed.
No
reproductive
or
developmental
effects
were
observed
at
doses
up
to
approximately
116
mg/
kg­
day
in
females
or
68
mg/
kg­
day
in
males
in
studies
conducted
in
Sprague­
Dawley
rats
(
NTP,
1998).
Adverse
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VII
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7
reproductive
or
developmental
effect
were
not
observed
in
rabbits
exposed
to
doses
as
high
as
55
or
76
mg/
kg­
day
in
drinking
water
on
gestation
days
6
to
29
(
CCC,
2000c,
d).
Increased
incidences
of
sternebral
aberrations
(
Ruddick
et.
al.,
1983)
and
decreased
ossification
sites
in
the
forelimb
and
hindlimb
(
CCC,
2000b)
have
been
observed
in
Sprague­
Dawley
rats
administered
bromodichloromethane
in
corn
oil
and
drinking
water,
respectively,
at
doses
which
induced
maternal
toxicity.

Reproductive
effects
of
bromodichloromethane
have
been
noted
in
rodent
assays.
Klinefelter
et
al.
(
1995)
observed
effects
on
sperm
motility
in
rats
administered
39
mg/
kg­
day
in
drinking
water
for
52
weeks,
but
these
effects
were
not
accompanied
by
histopathological
changes
in
male
reproductive
tissues.
Narotsky
et
al.
(
1997)
observed
a
significantly
increased
incidence
of
full
litter
resorption
(
FLR)
in
F344
rats
treated
with
75
mg/
kg­
day
bromodichloromethane
by
aqueous
gavage
throughout
the
period
of
organogenesis.
This
effect
was
described
as
an
all­
or­
nothing
phenomenon,
in
that
the
litter
was
either
fully
resorbed
or
appeared
normal
at
parturition.
This
pattern
was
interpreted
by
the
study
authors
as
evidence
for
a
maternally­
mediated
mechanism,
rather
than
a
direct
effect
of
bromodichloromethane
on
the
developing
embryo.
Bielmeier
et
al.
(
2001)
observed
increased
incidence
of
FLR
in
F344
rats
treated
with
75
or
100
mg/
kg­
day
BDCM
by
aqueous
gavage
on
one
or
more
days
during
the
interval
from
gestation
day
6
to
10.
This
response
was
strain­
specific
and
FLR
was
not
observed
in
Sprague­
Dawley
rats
treated
with
up
to
100
mg/
kg­
day
on
gestation
days
6
to10.

Data
from
Epidemiological
Studies
There
is
evidence
from
epidemiological
studies
which
suggests
that
exposure
to
bromodichloromethane
is
possibly
associated
with
reproductive
effects.
An
epidemiological
study
by
Waller
et
al.
(
1998)
found
an
association
between
consumption
of
trihalomethanes
in
drinking
water
and
increased
risk
of
spontaneous
abortion.
Analysis
of
personal
exposure
to
individual
trihalomethanes
demonstrated
that
bromodichloromethane
had
the
strongest
association
to
with
spontaneous
abortion.
When
all
four
individual
trihalomethanes
(
including
chloroform)
were
simultaneously
included
as
variables
in
a
logistic
regression
model,
high
personal
bromodichloromethane
exposure
had
an
odds
ratio
of
3.0
for
spontaneous
abortion
(
95%
C.
I.
1.4,
6.6).
King
et
al.
(
2000)
examined
the
relationship
between
stillbirth
of
fetuses
weighing
more
than
500
g
and
chlorination
byproducts
in
drinking
water.
Increased
risk
of
stillbirth
was
associated
with
total
trihalomethane
concentration,
chloroform
concentration,
and
bromodichloromethane
concentration.
The
strongest
association
was
observed
for
bromodichloromethane,
where
risk
doubled
in
women
exposed
to
concentrations
of
20
µ
g/
L
or
more
when
compared
to
women
consuming
water
containing
concentrations
less
than
5
µ
g/
L.

Both
population­
based
studies
had
reasonable
cohorts,
recruitment
procedures,
and
screening
protocols.
The
number
of
subjects
was
adequate,
and
the
information
regarding
fetal
losses
was
credible.
Therefore,
the
conclusion
that
increased
fetal
loss
in
humans
occurs
in
association
with
the
exposure
measurements
that
were
made
is
quite
convincing.
A
weakness
of
both
population
based
studies,
however,
is
that
the
measurement
of
the
trihalomethanes
appeared
to
be
at
the
source
or
at
intermediate
sources
rather
than
at
the
tap.
Thus,
the
Draft
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20,
2003
VII
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bromodichloromethane
concentrations
that
were
measured
could
be
different
than
the
actual
exposures
(
levels
at
the
tap)
and/
or
the
bromodichloromethane
could
be
a
surrogate
for
the
actual
toxicant
in
both
population­
based
studies.
Although
these
human
studies
raise
significant
concern,
the
occurrence
of
multiple
disinfection
byproducts
in
drinking
water
prevents
the
conclusion
that
bromodichloromethane
was
the
specific
causative
agent
for
the
observed
effects.

Full
Litter
Resorption
in
F344
Rats
As
indicated
above,
two
oral
exposure
studies
of
bromodichloromethane
have
observed
increased
incidences
of
full
litter
resorption
(
FLR)
in
the
F344
strain
of
rat.
Narotsky
et
al.
(
1997)
observed
a
significantly
increased
incidence
of
full
litter
resorption
(
FLR)
in
F344
rats
treated
with
75
mg/
kg­
day
bromodichloromethane
by
aqueous
gavage
throughout
the
period
of
organogenesis.
These
studies
are
of
particular
interest
because
of
epidemiological
association
between
bromodichloromethane
ingestion
and
increased
risk
of
spontaneous
abortion.

In
light
of
human
epidemiological
studies
possibly
implicating
bromodichloromethane
in
pregnancy
loss,
Bielmeier
et
al.
(
2001)
conducted
a
series
of
experiments
to
further
elucidate
the
mode
of
action
of
this
compound
on
full
litter
resorption
in
F344
rats.
Their
experiments
investigated
the
effects
of
bromodichloromethane
on
serum
levels
of
progesterone
(
necessary
for
the
maintenance
of
pregnancy)
and
luteinizing
hormone
(
LH),
which
helps
maintain
the
corpus
luteum
and
luteal
function
(
the
secretion
of
progesterone).
In
the
rat,
secretion
of
progesterone
by
the
corpus
luteum
is
required
up
to
GD
17
for
maintenance
of
pregnancy.
Luteal
function
is
established
and
maintained
in
the
rat
through
a
series
of
mechanisms
that
require
cervical
stimulation
by
coitus,
a
pituitary
response
to
this
cervical
stimulation,
and
a
secretion
of
hormones
by
the
products
of
conception.
In
the
newly
formed
corpus
luteum,
progesterone
secretion
is
autonomous.
However,
continued
function
requires
luteotropic
hormones
secreted
by
the
pituitary.
In
many
species
LH
is
considered
to
be
the
most
important
luteotropic
factor.
However,
in
rats
prolactin
has
been
identified
as
the
primary
luteotropic
factor.

The
hormonal
requirements
of
the
corpus
luteum
change
with
the
physiological
stage
of
the
pregnancy.
Early
in
the
rat
pregnancy,
prolactin
secreted
by
the
pituitary
is
essential
for
luteal
function.
Starting
at
about
GD
6,
LH
also
is
required
for
maintenance
of
luteal
function
in
addition
to
prolactin.
The
corpora
lutea
are
dependent
on
LH
until
GD
11,
at
which
time
luteal
maintenance
becomes
dependent
on
one
or
more
placental
prolactin­
like
hormones
(
prolactogens)
(
Gibori
et
al.,
1988).

Bielmeier
et
al.
(
2001)
investigated
the
effects
of
bromodichloromethane
in
rats
during
two
physiologically
different
stages
of
pregnancy:
GD
6
to
10,
which
encompasses
the
LHdependent
stage
of
pregnancy
and
GD
11
to
15,
which
is
an
LH­
independent
stage.
These
authors
observed
increased
incidences
of
FLR
in
F344
rats
treated
with
75
or
100
mg/
kg­
day
BDCM
on
one
or
more
days
during
the
LH­
dependent
interval
from
gestation
day
6
to
10,
but
not
in
rats
treated
during
the
LH­
independent
period.
Thus,
the
critical
period
for
induction
of
FLR
was
limited
to
the
luteinizing
hormone
(
LH)­
dependent
phase
of
pregnancy,
suggesting
that
bromodichloromethane
may
disrupt
pregnancy
via
a
LH­
mediated
mode
of
action.
The
treatment
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VII
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period
also
overlapped
the
period
when
maintenance
of
pregnancy
is
dependent
on
progesterone
produced
by
the
ovary.
The
response
to
bromodichloromethane
was
strain­
specific
and
FLR
was
not
observed
in
Sprague­
Dawley
rats
treated
with
up
to
100
mg/
kg­
day
on
GD
6
to10.
Measurement
of
serum
LH
and
progesterone
levels
indicated
that
FLR
was
accompanied
by
a
marked
reduction
in
progesterone
concentration
without
a
corresponding
drop
in
LH
levels.
The
failure
of
bromodichloromethane
to
exert
adverse
effects
after
the
LH­
dependent
window,
the
reduction
in
serum
progesterone
level,
and
the
unchanged
serum
luteinizing
hormone
levels
led
this
group
to
conclude
that
the
target
of
toxicity
was
the
ovary
and
that
the
mode
of
action
was
a
reduced
sensitivity
of
the
corpus
luteum
to
luteinizing
hormone.

Bielmeier
et
al.
(
2001)
have
emphasized
that
there
are
significant
differences
between
rats
and
humans
in
the
hormonal
maintenance
of
pregnancy.
However,
these
authors
have
noted
that
rats
and
humans
are
similar
in
that
either
LH
or
human
chorionic
gonadotropin
(
hCG),
which
act
via
the
same
receptor,
are
required
during
specific
gestational
periods
to
maintain
pregnancy.
In
the
opinion
of
these
authors,
the
findings
in
rats
may
be
relevant
to
observations
of
adverse
pregnancy
outcomes
in
human
epidemiological
studies
if
bromodichloromethane
disrupts
pregnancy
by
diminishing
luteal
responsiveness
to
LH
via
an
effect
on
the
LH/
hCG
receptor,.
However,
additional
research
is
required
to
determine
whether
the
findings
in
rats
are
potentially
relevant
to
humans.

As
noted
by
Narotsky
et
al.
(
1997),
it
is
important
to
recognize
that
bromodichloromethane­
induced
FLR
occurs
in
F344
rats
at
doses
several
orders
of
magnitude
greater
than
would
be
experienced
by
humans
consuming
highly
contaminated
water.
For
example,
a
dose
of
75
mg/
kg­
day
is
approximately
15,000­
fold
higher
than
human
intake
assuming
a
body
weight
of
70
kg,
a
bromodichloromethane
concentration
of
180
µ
g/
L,
and
water
consumption
of
2L/
day.
However,
the
possibility
that
humans
are
more
sensitive
to
bromodichloromethane
than
rats
cannot
be
dismissed.
At
this
time,
data
are
insufficient
to
evaluate
this
possibility.

b.
Childhood
Cancer
and
Noncancer
Effects
Bioactivation
to
reactive
metabolites
is
an
apparent
prerequisite
for
toxicity
and
carcinogenicity
of
the
brominated
trihalomethanes.
Therefore,
an
important
issue
in
the
assessment
of
childhood
risk
of
cancer
and
other
adverse
effects
is
whether
the
enzymes
responsible
for
metabolism
are
more
active
in
fetuses,
neonates,
and
or
children
than
in
adults.
This
section
evaluates
the
available
data
for
developmental
expression
and/
or
activity
of
three
key
metabolizing
enzymes
that
are
known
or
anticipated
to
bioactivate
the
brominated
trihalomethanes:
CYP2E1,
CYP2B1/
2
(
in
rodents
only),
and
glutathione­
S­
transferase
theta.

CYP2E1
Carcinogenicity
of
brominated
trihalomethanes
has
been
shown
to
be
at
least
partly
related
to
bioactivation
by
the
cytochrome
P450
isoform
CYP2E1
(
U.
S.
EPA,
1994b).
Thus,
a
higher
level
of
CYP2E1
activity
in
children
relative
to
adults
might
predispose
children
to
greater
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2003
VII
­
10
toxicity.
Studies
of
human
fetal
liver
have
produced
contradictory
results,
but
suggest
that
CYP2E1
protein
is
either
not
expressed
or
is
expressed
at
levels
lower
than
in
adults
(
Hakkola
et
al.,
1998).
Carpenter
et
al.
(
1996)
detected
immunoreactive
CYP2E1
protein
in
liver
samples
from
fetuses
ranging
from
16
to
24
weeks
in
gestational
age.
The
samples
obtained
were
from
fetuses
whose
mothers
did
not
have
a
history
of
alcohol
use.
The
immunoreactive
protein
exhibited
a
slightly
lower
molecular
weight
than
observed
for
CYP2E1
from
adult
liver
samples.
Expression
of
the
corresponding
mRNA
was
confirmed
in
a
fetal
liver
sample
of
19
weeks
gestational
age
by
reverse
transcriptase­
polymerase
chain
reaction
(
RT­
PCR).
However,
CYP2E1
mRNA
was
not
detectable
in
a
fetal
liver
sample
of
10
weeks
gestational
age,
suggesting
(
in
the
opinion
of
the
study
authors)
that
CYP2E1
expression
may
be
related
to
specific
stages
of
fetal
development.
The
catalytic
capability
of
CYP2E1
protein
in
human
fetal
microsomes
was
demonstrated
by
measuring
the
rate
of
ethanol
oxidation
to
acetaldehyde.
The
rate
of
conversion
varied
from
12
to
27%
of
that
measured
in
adult
microsomes.
Treatment
of
fetal
hepatocytes
in
primary
culture
with
ethanol
or
clofibrate
indicated
that
fetal
CYP2E1
protein
is
inducible
(
approximately
two­
fold
compared
to
untreated
cells).

Viera
et
al.
(
1996)
detected
small
amounts
of
CYP2E1
mRNA
in
fetal
liver
samples
(
approximately
5
to
10%
of
the
levels
in
adult
liver)
collected
from
fetuses
aged
14
to
40
weeks.
However,
these
authors
could
not
detect
immunoreactive
CYP2E1
protein
in
any
of
27
fetal
liver
samples
Other
studies
have
failed
to
detect
either
CYP2E1
protein
or
mRNA
in
fetal
liver
samples.
Cresteil
et
al.
(
1985)
and
Komori
et
al.
(
1989)
did
not
detect
immunoreactive
protein
or
mRNA
in
fetal
liver
samples
of
less
than
16
weeks
gestational
age.
Jones
et
al.
(
1992)
did
not
detect
CYP2E1
mRNA
or
protein
in
liver
samples
that
were
of
similar
gestational
age
(
16
to
18
weeks)
to
the
samples
examined
by
Carpenter
et
al.
(
1996).
Juchau
and
Yang
(
1996)
did
not
detect
CYP2E1
mRNA
by
RT­
PCR
in
human
embryonic
tissues
between
days
45
and
60
of
gestation.
The
factors
contributing
to
the
different
results
are
unknown,
but
may
include
interindividual
variability,
gestational
age
of
the
tissue
examined
(
for
the
samples
less
than
16
weeks
gestational
age),
or
the
existence
of
factors
other
than
developmental
stage
that
control
expression.

Information
on
the
presence
of
CYP2E1
in
human
fetal
tissues
other
than
the
liver
is
limited.
Viera
et
al.
(
1998)
examined
the
mRNA
content
of
human
fetal
lung
and
kidney.
CYP2E1
mRNA
was
expressed
at
a
very
low
level
in
both
tissues
and
the
levels
remained
stable
after
birth.
Studies
of
human
fetal
brain
tissue
indicate
that
CYP2E1
is
expressed
in
human
embryonic
brain
tissue
(
see
Juchau
et
al.,
1994)
and
that
relatively
low
levels
of
CYP2E1
mRNA,
immunoreactive
protein,
and
catalytically
active
protein
are
present
during
the
early
fetal
period
of
development
(
Brzezinski
et
al.,
1999).
In
one
study,
a
dramatic
increase
in
CYP2E1
was
observed
at
approximately
gestation
day
50,
and
a
fairly
constant
level
was
maintained
until
at
least
day
113
(
Brzezinski
et
al.,
1999).
The
relevance
of
the
data
for
lung
and
brain
is
uncertain,
since
these
organs
are
not
known
to
be
targets
for
brominated
trihalomethane
toxicity.

Viera
et
al.
(
1996)
investigated
age­
related
variations
in
human
CYP2E1
protein
levels
and
catalytic
activity
from
birth
through
adulthood.
These
authors
observed
a
rapid
increase
in
the
immunoreactive
CYP2E1
microsomal
content
within
24
hours
after
birth
that
was
independent
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February
20,
2003
VII
­
11
of
the
gestational
age
of
the
newborn.
This
activation
was
accompanied
by
a
demethylation
of
cytosine
residues
in
the
5'­
regulatory
region
of
the
gene,
suggesting
that
methylation
of
specific
residues
prevents
transcription
in
the
fetal
liver.
The
CYP2E1
protein
level
gradually
increased
during
the
first
year
and
reached
the
adult
level
in
children
aged
1
to
10
years.
CYP2E1
catalytic
activity
was
assessed
by
determination
of
in
vitro
hydroxylation
of
chlorzoxazone
in
89
microsomal
preparations.
Chlorzoxazone
hydroxylation
activity
increased
within
24
hours
after
birth
and
steadily
increased
during
the
first
year.
Catalytic
activity
reached
adult
levels
at
age
1
to
10
years.

Animal
studies
of
CYP2E1
expression
during
development
have
given
variable
results.
One
study
indicated
that
CYP2E1
is
expressed
in
the
fetal
rat
liver
and
placenta
and
that
levels
are
increased
in
rat
pups
exposed
to
ethanol
in
utero
or
via
lactation
(
Carpenter
et
al.,
1997).
Liver
samples
from
rat
fetuses
exposed
to
ethanol
in
utero
showed
a
2.4­
fold
increase
in
protein
levels
and
1.5­
fold
increase
in
catalytic
activity
(
Carpenter
et
al.,
1997).
Other
authors
have
reported
that
hepatic
CYP2E1
gene
transcription
in
rats
is
activated
at
birth
and
that
the
amount
of
CYP2E1
reaches
a
peak
prior
to
weaning
(
see
Ronis
et
al.,
1996).
The
protein
level
then
falls
to
approximately
25%
of
the
peak
level
and
remains
stable
into
adulthood
(
Ronis
et
al.,
1996).

The
regulation
of
CYP2E1
is
complex
when
examined
at
both
the
molecular
(
Lieber
1997)
and
physiological
(
Ronis
et
al.
1996)
levels.
The
factors
and
processes
responsible
for
the
increase
in
CYP2E1
protein
levels
and
activity
at
birth
have
not
been
clearly
identified.
At
the
physiological
level,
there
is
some
evidence
from
rodent
studies
to
suggest
that
growth
hormone
regulates
the
constitutive
expression
of
CYP2E1
(
Ronis
et
al.,
1996).
The
reduction
of
CYP2E1
from
peak
levels
before
weaning
is
reported
to
coincide
with
the
increased
levels
of
growth
hormone
and
with
development
of
adult
levels
of
growth
hormone
receptors
(
Ronis
et
al.,
1996).
The
occurrence
of
peak
expression
after
birth
has
been
attributed
to
a
role
of
CYP2E1
in
gluconeogenesis,
since
there
is
a
very
high
demand
for
energy
production
from
glucose
at
this
developmental
stage
(
see
Ronis
et
al.
1996;
Viera
et
al.
1996).

CYP2B1/
2
(
Rodents)

Research
conducted
by
Pankow
et
al.
(
1997)
suggests
that
the
closely­
related
CYP
isoforms
2B1
and
2B2
participate
in
the
catabolism
of
dibromochloromethane
in
rats.
These
isoforms
show
greater
than
97%
homology
of
amino
acid
sequence
and
have
highly
similar
genomic
organization.
To
date,
these
isoforms
have
not
been
reported
in
adult
or
fetal
human
tissues
(
Nelson
et
al.,
1996;
Juchau
et
al.,
1998).
Omiecinski
et
al.
(
1990)
detected
low
levels
of
CYP2B
isoform
mRNA
in
fetal
rat
liver
on
gestation
day
15
(
the
earliest
day
in
development
when
the
authors
were
able
to
macroscopically
recognize
and
dissect
the
fetal
liver)
using
the
polymerase
chain
reaction
(
PCR).
Although
the
levels
of
mRNA
expression
were
"
substantially
lower"
lower
at
day
15
than
observed
later
in
development,
expression
was
clearly
inducible
by
pretreatment
of
pregnant
rats
with
phenobarbital.
Both
constitutive
and
phenobarbital­
induced
levels
of
mRNA
increased
with
developmental
age,
reaching
maximal
levels
at
approximately
three
weeks
postpartum.
No
measurements
of
CYP2B
activity
were
made
in
this
study,
so
it
is
not
known
whether
changes
in
mRNA
levels
were
paralleled
by
changes
in
catalytic
activity.
Draft
­
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February
20,
2003
VII
­
12
Juchau
et
al.
(
1998)
reviewed
a
series
of
experiments
that
employed
the
selective
substrate
probe
pentoxyresorufin
to
test
for
CYP2B1/
2
catalytic
activity
in
fetal
rat
tissues.
The
overall
conclusion
upon
examination
of
all
results
was
that
if
CYP2B
isoforms
are
expressed
in
fetal
rats,
they
occur
at
biologically
insignificant
levels.
Asoh
et
al.
(
1999)
examined
the
catalytic
activity
of
CYP2B
isoforms
in
fetal
rat
liver
and
found
very
low
activity,
a
finding
consistent
with
the
conclusion
of
Juchau
et
al.
(
1998).

Gebremichael
et
al.
(
1995)
investigated
the
postnatal
developmental
profile
of
CYP2B1
in
Sprague­
Dawley
rats.
CYP2B1
activity
was
detectable
as
early
as
seven
days
postnatally
and
exhibited
a
variable
pattern
of
expression
(
no
clear
trend
evident)
when
assayed
at
Days
14,
21,
50,
and
100.
Asoh
et
al.
(
1999)
examined
the
induction
of
CYP2B
isoforms
in
neonatal
rats.
The
level
of
CYP2B
catalytic
activity
was
markedly
higher
at
five
days
after
birth
relative
to
levels
observed
in
fetal
hepatic
tissue.
Oral
or
intraperitoneal
administration
of
phenobarbital
to
pregnant
rats
increased
the
level
of
CYP2B
expression
and
activity
in
neonates.
Overall,
these
findings
suggest
that
CYP2B
isoform
activity
is
likely
to
be
lower
in
fetuses
than
in
neonates
or
adults
and
that
increased
levels
of
activity
may
be
observed
in
fetuses
and
neonates
exposed
to
inducing
xenobiotics.
The
significance
of
this
information
for
risk
of
cancer
in
human
fetuses,
neonates,
and
children
is
uncertain
since,
as
noted
above,
the
CYP2B1/
2
isoforms
have
not
been
identified
in
humans.

Glutathione
S­
Transferase
Theta
Genotoxicity
studies
in
genetically
engineered
bacteria
indicate
that
brominated
trihalomethanes
can
also
be
activated
to
mutagens
by
the
product
of
the
glutathione
S­
transferase
(
GST)
theta
gene
GSTT1­
1
(
DeMarini
et
al.,
1997;
Landi
et
al.,
1999).
Children
and
the
fetus
could
potentially
experience
increased
risk
of
adverse
effects
if
the
activity
of
this
enzyme
was
higher
at
these
life
stages
than
in
adults.
Information
on
the
developmental
expression
of
GST
genes
is
currently
limited.
Although
other
classes
of
GSTs
(
alpha,
mu,
and
pi)
are
expressed
in
fetal
liver,
Mera
et
al.
(
1994)
reported
that
theta­
class
GSTs
were
expressed
in
only
adult
liver.
This
finding
suggests
that
the
fetus
does
not
experience
increased
risk
as
a
result
of
GST
thetamediated
mutagenicity.
The
occurrence
of
increased
risk
in
children
cannot
be
evaluated,
since
at
present
the
age
at
which
synthesis
of
the
GST
theta
is
induced
is
unknown.

c.
Childhood
Cancers:
Other
Considerations
Examination
of
childhood
cancer
data
compiled
by
the
National
Cancer
Institute
(
Ries
et
al.
1999)
indicates
that
the
incidence
of
hepatic,
renal,
and
intestinal
cancer
(
the
types
of
cancer
observed
in
animal
studies
of
brominated
trihalomethane
cancer
potential)
from
causes
other
than
genetic
predisposition
are
low.
Primary
neoplasms
of
the
liver
are
rare
in
children
younger
than
15
years
of
age.
There
has
been
little
change
in
the
liver
cancer
incidence
(
total
hepatic
tumors)
in
this
age
group
over
the
last
21
years,
with
rates
between
1.4
and
1.7
cases
per
million
throughout
the
time
period.
The
incidence
of
hepatocellular
carcinoma
(
the
type
of
neoplasm
observed
in
mice
treated
with
bromodichloromethane
or
dibromochloromethane)
decreased
in
children
younger
than
15
years
of
age
during
the
period
1975
to
1995.
The
incidence
rate
of
renal
Draft
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February
20,
2003
VII
­
13
carcinoma
remained
very
low
(
well
under
one
case
per
million)
in
children
younger
than
15
years
during
the
period
1975
to
1995.
Trend
data
were
not
available
in
Ries
et
al.
(
1999)
for
intestinal
cancer.
These
observations
are
consistent
with
(
but
do
not
prove)
a
conclusion
that
children
are
not
more
susceptible
than
adults.

d.
Conclusion
The
available
evidence
for
developmental
expression
of
enzymes
known
to
metabolize
brominated
trihalomethanes
supports
the
conclusion
that
children
do
not
experience
greater
risk
from
exposure
to
these
compounds
than
do
adults.
At
present
there
are
no
cancer
incidence
data
from
humans
to
suggest
that
brominated
trihalomethanes
contribute
to
increased
risk
of
cancer
in
children.

3.
Other
Potentially
Susceptible
Populations
a.
Subpopulations
with
altered
levels
of
CYP2E1
CYP2E1
catalyzes
the
metabolism
of
brominated
trihalomethanes
to
reactive
intermediates
that
mediate
toxicity.
Individuals
with
higher
levels
of
CYP2E1
activity
may
therefore
be
at
greater
risk
for
adverse
health
effects.
This
section
describes
factors
associated
with
increased
levels
of
CYP2E1
activity
and
subpopulations
who
may
be
at
increased
risk
as
a
result
of
these
factors.

Genetic
Polymorphisms
Significant
inter­
ethnic
differences
exist
in
CYP2E1
polymorphism
(
Ronis
et
al.,
1996;
Lieber
1997))
and
it
is
possible
that
these
differences
could
influence
susceptibility
to
toxic
effects.
The
CYP2E1
polymorphisms
currently
reported
in
the
literature
are
located
in
the
5'­
flanking
(
noncoding)
regions
of
the
gene,
while
the
coding
regions
of
the
gene
which
specify
sequence
appear
to
be
well
conserved
among
various
ethnic
groups
(
Ronis
et
al.
1996).
Mutations
in
the
5'­
region
of
a
gene
can
affect
the
regulation
gene
expression.
The
rare
mutant
c2
polymorphism
of
CYP2E1
is
reported
to
be
associated
with
higher
transcriptional
activity,
protein
levels,
and
catalytic
activity
than
the
more
common
wild
type
allele
(
Lieber
et
al.
1997).
As
reported
by
Lieber
et
al.
(
1997),
the
highest
frequency
of
the
c2
allele
occurs
in
the
Taiwanese
(
0.28)
and
Japanese
(
0.19
to
0.27)
populations.
The
frequencies
in
African­
Americans,
European­
Americans,
and
Scandinavians
are
much
lower,
generally
in
the
range
0.01
and
0.05.
Efforts
to
link
the
occurrence
of
the
c2
allele
to
higher
rates
of
CYP2E1­
mediated
liver
disease
have
yielded
inconsistent
results.
Thus,
the
functional
significance
of
CYP2E1
polymorphism
is
presently
uncertain,
and
no
conclusion
can
as
yet
be
drawn
about
the
relative
risk
for
different
ethnic
populations
exposed
to
brominated
trihalomethanes.

Altered
Physiological
or
Health
States
Draft
­
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or
Quote
February
20,
2003
VII
­
14
The
physiological
functions
of
CYP2E1
include
lipid
metabolism
and
ketone
utilization
(
Lieber,
1997).
Induction
of
CYP2E1
is
observed
in
many
conditions
that
elevate
circulating
levels
of
lipids,
including
consumption
of
a
high­
fat
or
low­
carbohydrate
diet,
starvation,
obesity,
and
insulin­
dependent
diabetes.
Among
the
individuals
likely
to
be
affected
by
such
conditions,
diabetics
constitute
the
most
clearly­
defined
susceptible
population.
Induction
of
CYP2E1
in
uncontrolled
insulin­
dependent
diabetes
is
well­
studied.
In
animals,
this
induction
results
in
elevated
levels
of
CYP2E1
in
the
liver,
kidney,
and
lung
(
Ioannides
et
al.,
1996).
Acetone
(
a
substrate
of
CYP2E1)
is
thought
to
be
the
inducing
compound
(
Ronis
et
al.,
1996).
As
a
result
of
induction,
diabetic
animals
are
more
susceptible
to
the
toxicity
of
some
chemicals
metabolized
by
CYP2E1.
While
there
are
no
specific
data
for
the
brominated
trihalomethanes,
this
phenomenon
has
been
demonstrated
for
other
halogenated
compounds
including
chloroform,
carbon
tetrachloride,
trichloroethylene,
and
bromobenzene
(
Ioannides
et
al.,
1996).
Because
the
animal
and
human
orthologues
of
CYP2E1
show
similar
substrate
specificity
and
bioactivation
potential,
it
is
possible
that
diabetic
humans
may
also
be
more
susceptible
to
CYP2E1­
mediated
toxicity.
As
CYP2E1
levels
are
reduced
by
insulin
treatment,
increased
toxicity
would
be
anticipated
only
in
poorly
controlled
or
uncontrolled
diabetics
(
Ioannides
et
al.,
1996).

Alcohol
consumption
CYP2E1
contributes
to
the
metabolism
of
ethanol
in
humans
and
animals.
Consumption
of
ethanol
induces
CYP2E1
and
chronic
alcohol
consumption
is
reported
to
result
in
as
much
as
a
10­
fold
induction
(
Lieber,
1997).
Hence,
concurrent
exposure
to
ethanol
and
brominated
trihalomethanes
may
increase
susceptibility
to
adverse
health
effects.
This
interaction
is
of
concern
because
concurrent
exposure
to
brominated
trihalomethanes
and
ethanol
is
likely
to
occur
in
a
significant
number
of
people.
At
present,
there
are
no
human
or
animal
studies
which
examine
this
interaction
for
brominated
trihalomethanes.
However,
Wang
et
al.
(
1994)
reported
that
a
single
100
mg/
kg
oral
dose
of
ethanol
administered
to
rats
significantly
increased
the
toxicity
of
the
structurally­
related
trihalomethane
chloroform
(
also
metabolized
by
CYP2E1).
Lieber
(
1997)
noted
that
the
hepatotoxicity
of
commonly
used
industrial
solvents
(
e.
g.
carbon
tetrachloride,
bromobenzene,
and
vinylidene
chloride)
and
anesthetics
(
enflurane
and
halothane)
was
increased
in
heavy
drinkers,
with
a
pattern
of
damage
that
was
consistent
with
the
selective
expression
and
induction
of
CYP2E1
in
certain
regions
of
the
liver.

Concurrent
exposure
to
other
CYP2E1
inducers
including
pharmaceuticals
Because
CYP2E1
is
highly
inducible
by
a
wide
range
of
xenobiotic
compounds,
prior
exposure
to
such
inducers
may
potentially
play
a
significant
role
in
brominated
trihalomethane
toxicity.
Known
inducers
of
CYP2E1
include
certain
therapeutic
agents
(
acetaminophen,
isoniazid),
volatile
anaesthetics
(
halothane,
isoflurane),
and
solvents
(
acetone,
benzene,
carbon
tetrachloride,
trichloroethylene)
(
Raucy,
1995).
Individuals
exposed
to
or
consuming
these
inducers
on
a
regular
basis
may
therefore
be
at
greater
risk
for
brominated
trihalomethane
toxicity.

b.
Subpopulations
with
altered
levels
of
glutathione
S­
transferase
theta
Draft
­
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Cite
or
Quote
February
20,
2003
VII
­
15
Individuals
with
Genetic
Polymorphisms
Genotoxicity
studies
in
bacteria
(
discussed
in
section
V.
F.
1)
indicate
that
brominated
trihalomethanes
can
be
activated
to
mutagens
by
the
product
of
the
glutathione
S­
transferase
theta
gene
GSTT1­
1
(
DeMarini
et
al.,
1997;
Landi
et
al.,
1999).
If
similar
pathways
for
bioactivation
exist
in
humans,
GSTT1­
1
polymorphism
may
influence
susceptibility
to
brominated
trihalomethane­
mediated
toxicity.
GSTT1­
1
is
characterized
by
a
deletion
polymorphism
which
results
in
total
loss
of
glutathione
S­
transferase­
 
activity
in
individuals
(
10
to
60%
of
the
population
depending
upon
ethnicity
and
race)
homozygous
for
the
null
genotype
(
GSTT1­
1­/­).
Individuals
who
are
heterozygous
for
GSTT­
1
(
GSTT1­
1+/­)
have
intermediate
levels
of
enzyme
activity,
while
individuals
homozygous
for
GSTT­
1
(
GSTT1­
1+/+)
have
the
highest
levels.
Landi
et
al.
(
1999)
have
suggested
that
GSTT1­
1+/+
individuals
may
experience
excess
genotoxic
risk
when
exposed
to
brominated
trihalomethanes,
particularly
in
organs
which
express
glutathione­
Stransferase
theta
and
come
in
direct
contact
with
brominated
trihalomethanes.
Potential
target
sites
would
include
the
gastrointestinal
tract
and
the
bladder.

Concurrent
Exposure
to
Inducers
If
GSTT­
1­
mediated
pathways
for
bioactivation
of
brominated
trihalomethanes
exist
in
humans,
factors
which
induce
this
enzyme
may
increase
the
risk
of
adverse
health
effects
from
exposure.
Although
GSTT­
1
is
constitutively
expressed,
the
level
of
its
expression
can
be
altered
by
exposure
to
exogenous
chemicals.
Landi
(
2000)
has
summarized
information
on
factors
which
increase
expression
of
the
enzyme.
In
rats,
aspirin
increased
GSTT­
1
levels
in
the
colon.
Alphatocopherol
coumarin;
and
other
anticarcinogenic
drugs
increased
gastric
and
esophageal
levels;
and
indole­
3­
carbinol
and
coumarin
increased
GSTT­
1
levels
in
the
liver.
In
mice,
phenobarbital
induced
hepatic
GSTT­
1
levels.
Data
for
humans
are
limited,
but
there
are
indications
that
the
dietary
intake
of
cruciferous
vegetables
enhances
the
expression
of
GSTT­
1.
It
is
possible
that
consumption
of
these
substances
by
GSTT­
1
positive
individuals
could
result
in
increased
risk
of
adverse
effects.
However,
there
are
presently
no
data
available
for
evaluation
of
this
hypothesis.

c.
Subpopulations
with
altered
levels
of
putative
protective
compounds
Glutathione
depletion
has
been
observed
to
increase
the
hepatotoxicity
of
bromodichloromethane
(
Gao
et
al.,
1996).
On
the
basis
of
these
data,
Gao
et
al.
(
1996)
proposed
that
populations
with
low
baseline
levels
of
glutathione
(
e.
g.,
due
to
dietary
deficiencies
of
glutathione
precursors
such
as
cysteine
and
selenium)
may
be
more
sensitive
to
bromodichloromethane­
induced
toxicity.

d.
Possible
gender
differences
Apparent
gender­
related
differences
in
the
toxicity
of
brominated
trihalomethanes
have
been
noted
in
studies
where
male
and
female
animals
were
exposed
concurrently
(
e.
g.
Aida,
1992a;
Daniel,
1990;
NTP,
1987,
1989a;
Tobe
et
al.,
1982a).
In
general,
male
rats
and
mice
Draft
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February
20,
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VII
­
16
appear
to
be
somewhat
more
sensitive
to
the
hepatic
and
renal
toxicity
induced
by
brominated
trihalomethanes
than
are
females,
although
there
are
exceptions
to
this
pattern
(
eg.
the
chronic
oral
exposure
study
of
bromoform
conducted
by
NTP,
1989a
in
mice
and
the
short­
term
study
of
bromoform
conducted
by
Aida
et
al.,
1992a).
While
the
basis
for
the
apparent
greater
sensitivity
of
males
is
unknown,
the
difference
may
be
related
to
gender­
specific
differences
in
the
level
of
enzymes
responsible
for
bioactivation
of
brominated
trihalomethanes
to
toxic
metabolites,
or
to
gender­
specific
differences
in
cellular
protective
mechanisms.
It
is
important
to
note
that
at
present
there
is
no
evidence
for
gender­
related
differences
in
the
activity
levels
of
CYP2E1
or
GSTT­
1
in
humans.

E.
Summary
It
is
generally
believed
that
the
toxicity
of
the
brominated
trihalomethanes
is
related
to
their
metabolism.
This
conclusion
is
based
largely
on
the
observation
that
liver
and
kidney,
the
chief
target
tissues
for
these
compounds,
are
also
the
primary
sites
of
their
metabolism.
In
addition,
treatments
which
increase
or
decrease
metabolism
also
tend
to
increase
or
decrease
trihalomethane­
induced
toxicity
in
parallel.

Metabolism
of
brominated
trihalomethanes
is
believed
to
occur
via
oxidative
and
reductive
pathways.
Limited
structure­
activity
data
for
brominated
trihalomethanes
and
the
structurallyrelated
trihalomethane
chloroform
suggest
that
bromination
may
influence
the
proportion
of
compound
metabolized
via
the
oxidative
and
reductive
pathways,
with
brominated
compounds
being
more
extensively
metabolized
by
the
reductive
pathway.
Additional
evidence
suggests
that
a
GSH­
mediated
pathway
may
play
an
important
role
in
metabolism
of
brominated
trihalomethanes.
These
data
raise
the
possibility
that
brominated
trihalomethanes
may
induce
adverse
effects
(
toxicity
and
carcinogenicity)
via
several
different
pathways.

The
precise
biochemical
mechanisms
which
link
brominated
trihalomethane
metabolism
to
toxicity
have
not
been
characterized,
but
many
researchers
have
proposed
that
toxicity
results
from
the
production
of
reactive
intermediates.
Reactive
intermediates
may
arise
from
either
the
oxidative
(
dihalocarbonyls)
or
the
reductive
(
free
radicals)
pathways
of
metabolism.
Such
reactive
intermediates
are
known
to
form
covalent
adducts
with
various
cellular
molecules,
and
may
impair
the
function
of
those
molecules
and
cause
cell
injury.
Free
radical
production
may
also
lead
to
cell
injury
by
inducing
lipid
peroxidation
in
cellular
membranes.
Direct
evidence
showing
a
relationship
between
the
level
of
covalent
binding
intermediates
generated
by
either
pathway
and
the
extent
of
toxicity
is
not
available
for
the
brominated
trihalomethanes.
Manipulation
of
cellular
glutathione
levels
suggests
that
this
compound
may
play
a
protective
role
in
brominated
trihalomethane­
induced
toxicity.

Individual
brominated
trihalomethanes
have
been
shown
to
induce
tumors
in
laboratory
animals.
The
mechanism
by
which
brominated
trihalomethanes
induce
tumors
in
target
tissues
has
not
been
fully
characterized.
DNA
adducts
can
be
formed
by
interaction
of
reactive
metabolites
(
resulting
from
oxidative
and
reductive
metabolism)
with
DNA.
In
addition,
preliminary
evidence
suggested
that
DNA
adducts
can
be
formed
through
conjugation
with
glutathione
and
Draft
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February
20,
2003
VII
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17
bioactivation
of
the
resulting
conjugates.
Comparison
of
dose­
response
data
for
liver
toxicity
(
including
cell
proliferation)
and
tumorigenicity
in
mice
suggests
that
tumor
formation
occurs
at
concentrations
lower
than
those
which
stimulate
cell
proliferation.
No
evidence
for
increased
cell
proliferation
in
kidney
was
obtained
in
studies
using
doses
up
to
246
mg/
kg­
day
for
bromodichloromethane,
312
mg/
kg­
day
for
dibromochloromethane,
or
379
mg/
kg­
day
for
bromoform.

Interaction
with
agents
which
increase
or
decrease
the
activity
of
enzymes
responsible
for
metabolism
of
brominated
trihalomethanes
may
modify
carcinogenicity/
toxicity.
Pretreatment
with
inducers
of
CYP2E1
has
been
observed
to
increase
the
hepatotoxicity
of
bromodichloromethane
and
dibromochloromethane
in
male
rats.
Pretreatment
with
m­
xylene,
an
inducer
of
the
CYP2B1/
CYP2B2
isoforms,
increased
the
hepatotoxicity
of
dibromochloromethane
in
male
rats.
Conversely,
administration
of
the
cytochrome
P450
inhibitor
1­
aminobenzotriazole
prevented
bromodichloromethane­
induced
hepatotoxicity
in
rats.
Recent
findings
indicating
possible
glutathione­
mediated
metabolism
of
brominated
trihalomethanes
suggest
that
treatments
or
agents
which
alter
glutathione­
S­
transferase
activity
could
potentially
modify
the
toxicity
of
brominated
trihalomethanes.

The
severity
of
brominated
trihalomethane
toxicity
is
potentially
affected
by
the
vehicle
of
administration.
In
a
study
of
vehicle
effects
on
the
acute
toxicity
of
bromodichloromethane,
a
high
dose
(
400
mg/
kg)
of
the
chemical
was
more
hepato­
and
nephrotoxic
when
given
in
corn
oil
compared
to
aqueous
administration,
but
this
difference
was
not
evident
at
a
lower
dose
(
200
mg/
kg).

A
number
of
potentially
sensitive
subpopulations
have
been
identified
for
health
effects
of
brominated
trihalomethanes.
A
growing
body
of
scientific
evidence
indicates
that
children
may
suffer
disproportionately
from
some
environmental
health
risks.
These
risks
may
arise
because
the
neurological,
immunological,
and
digestive
systems
of
children
are
still
developing.
In
addition,
children
may
incur
greater
exposure
because
they
eat
more
food,
consume
more
fluids,
and
breathe
more
air
in
proportion
to
their
body
weight
when
compared
to
adults.
U.
S.
EPA
has
identified
three
key
questions
to
consider
when
evaluating
health
risks
to
children
from
drinking
water
disinfection
byproducts
(
DBP),
including
the
brominated
trihalomethanes:
1)
Is
there
information
which
shows
that
the
DBP
causes
effects
in
the
developing
fetus
or
impairs
ability
to
conceive
and
bear
children?
2)
If
the
DBP
causes
cancer,
are
children
more
likely
to
be
affected
by
it
than
are
adults?
and
3)
If
the
DBP
causes
a
noncancer
toxic
effect,
are
children
more
likely
to
be
affected
by
it
than
are
adults?
The
available
data
for
dibromochloromethane
and
bromoform
suggest
that
developmental
effects
in
animals
occur
only
at
doses
which
cause
maternal
toxicity
and
at
doses
lower
than
those
which
induce
histopathological
effects
in
the
liver
and
kidney.
There
is
no
evidence
that
these
compounds
impair
the
ability
to
conceive
or
have
children.
In
animal
studies,
exposure
to
bromodichloromethane
resulted
in
reduced
sperm
motility;
this
effect
was
not
accompanied
by
histopathologic
changes
in
the
male
reproductive
system.
Exposure
of
pregnant
F344
rats
during
gestation
days
6­
9
caused
full
litter
resorption.
This
response
was
not
observed
in
similarly
exposed
pregnant
Sprague­
Dawley
rats.
Epidemiological
studies
have
found
an
association
between
exposure
to
bromodichloromethane
in
drinking
water
and
increased
Draft
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February
20,
2003
VII
­
18
spontaneous
abortion
and
increased
stillbirth.
Although
these
studies
raise
concern
for
human
health
effects,
the
occurrence
of
multiple
disinfection
byproducts
in
drinking
water
prevents
the
conclusion
that
bromodichloromethane
is
the
causative
agent.

At
present,
there
are
no
cancer
data
which
indicate
that
brominated
trihalomethanes
contribute
to
increased
risk
of
cancer
in
children.
No
studies
were
located
which
examined
preor
post­
pubertal
cancer
rates
in
humans
in
relation
to
brominated
trihalomethane
exposure.
Cancer
bioassays
of
brominated
trihalomethanes
conducted
in
mice
and
rats
have
not
used
study
designs
that
included
perinatal
exposure.

The
available
evidence
suggests
that
the
toxic
effects
of
brominated
trihalomethanes
are
mediated
by
the
enzymes
CYP2E1,
CYP2B1/
2
(
in
rodents),
and
glutathione
S­
transferase
theta
(
GSTT­
1).
The
weight
of
evidence
from
studies
of
the
developmental
expression
of
these
enzymes
supports
the
conclusion
that
children
do
not
experience
greater
risk
from
brominated
trihalomethane
exposure
as
a
result
of
higher
metabolic
activity.

In
addition
to
children,
other
potentially
sensitive
populations
include
those
with
altered
levels
or
activity
of
CYP2E1
or
GSTT­
1
and
those
with
altered
levels
of
glutathione.
Factors
contributing
to
increases
in
CYP2E1
activity
potentially
include
genetic
polymorphisms;
altered
physiological
or
health
states;
alcohol
consumption;
and
concurrent
exposure
to
other
inducers,
including
some
pharmaceuticals
and
solvents.
Factors
contributing
to
increased
GSTT­
1
activity
include
genetic
polymorphisms
and
concurrent
exposure
to
inducers.
Based
on
observations
in
animals,
human
populations
with
reduced
levels
of
glutathione
as
a
result
of
dietary
deficiency
or
other
factors
may
experience
increased
sensitivity
to
the
toxic
effects
of
bromodichloromethane.

Apparent
gender­
related
differences
in
the
toxicity
of
brominated
trihalomethanes
have
been
noted
in
studies
where
male
and
female
animals
were
exposed
concurrently.
In
general,
male
rats
and
mice
appeared
to
be
more
sensitive
than
females
to
liver
and
renal
toxicity,
although
some
exceptions
to
this
pattern
have
been
noted.
There
is
no
evidence
for
a
similar
pattern
of
gender
response
in
humans.
Draft
­
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or
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20,
2003
VIII
­
1
VIII.
QUANTIFICATION
OF
TOXICOLOGICAL
EFFECTS
This
section
quantifies
the
toxicological
effects
of
brominated
trihalomethanes
based
on
health
effects
information
presented
in
Sections
V
and
VI.
At
present,
there
are
two
basic
approaches
to
quantification
of
toxicological
effects:
the
conventional
NOAEL/
LOAEL
approach
and
benchmark
dose
modeling.
Benchmark
dose
(
BMD)
modeling
(
U.
S.
EPA,
1995;
2000b)
was
chosen
as
the
preferred
approach
for
quantifying
toxicological
effects
of
the
brominated
trihalomethanes.
BMD
modeling
avoids
several
limitations
of
the
NOAEL/
LOAEL
approach,
including:
1)
the
slope
of
the
dose­
response
plays
little
role
in
determining
the
NOAEL;
2)
the
NOAEL
(
or
LOAEL)
is
limited
to
the
doses
tested
experimentally;
3)
the
determination
of
the
NOAEL
is
based
on
scientific
judgement,
and
is
subject
to
inconsistency;
and
4)
experiments
using
fewer
animals
tend
to
produce
larger
NOAELs,
and
as
a
result
may
produce
larger
health
advisories
(
HAs)
or
reference
doses
(
RfDs)
(
U.
S.
EPA,
1995)
that
may
not
be
sufficiently
protective
of
human
health.
In
contrast,
benchmark
doses
(
BMDs)
are
not
limited
to
the
experimental
doses,
appropriately
reflect
the
sample
size,
and
can
be
defined
in
a
statistically
consistent
manner.
The
BMD
approach
was
therefore
selected
for
quantification
of
toxicological
effects
of
the
brominated
trihalomethanes.
Values
for
HAs
and
RfDs
derived
using
the
conventional
NOAEL/
LOAEL
approach
are
presented
in
the
text
for
comparison
with
those
obtained
using
the
BMD
approach.

The
methods
employed
for
BMD
modeling
are
described
in
Appendix
A.
The
modeling
was
performed
using
the
BMDS
software
(
Version
1.2)
developed
by
the
U.
S.
EPA
National
Center
for
Environmental
Assessment.
The
BMDs
and
BMDLs
were
calculated
based
on
a
BMR
of
10%
extra
risk
for
all
quantal
endpoints
analyzed.
For
continuous
data,
the
BMR
was
defined
as
1.1
standard
deviations,
which
corresponds
to
an
additional
risk
of
approximately
10%
when
the
background
response
rate
is
assumed
to
be
1%
with
normal
variation
around
the
mean
and
constant
standard
deviation
(
Crump,
1995).
The
BMDL
10
was
defined
as
the
95%
lower
bound
on
the
corresponding
BMD
estimate.
Confidence
bounds
were
automatically
calculated
by
the
BMDS
software
using
a
likelihood
profile
method.

A.
Bromodichloromethane
1.
Noncarcinogenic
effects
a.
One­
day
Health
Advisory
Studies
of
the
acute
toxicity
of
bromodichloromethane
are
summarized
in
Table
VIII­
1.
Lilly
et
al.
(
1994)
administered
single
doses
of
bromodichloromethane
by
either
oil
or
aqueous
gavage
to
male
F344
rats
at
dose
levels
of
200
or
400
mg/
kg.
This
study
identified
a
LOAEL
of
200
mg/
kg­
day
based
on
histologic
lesions
in
the
kidney
and
changes
in
urinary
parameters.
A
NOAEL
value
was
not
identified
for
either
vehicle.
Data
for
hepatic
vacuolar
degeneration
and
renal
tubular
degeneration
obtained
using
the
aqueous
vehicle
were
modeled
using
the
BMDS
software.
BMD
and
BMDL
10
values
of
263
and
182
mg/
kg­
day,
respectively,
were
calculated
using
the
hepatic
data.
BMD
and
BMDL
10
values
of
131
and
8.9
mg/
kg­
day,
respectively,
were
Draft
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February
20,
2003
VIII
­
2
obtained
using
the
renal
data.
The
BMDL
10
for
renal
tubular
degeneration
is
the
lowest
calculated
across
studies,
but
is
not
considered
a
reliable
estimate
because
there
is
insufficient
information
to
accurately
characterize
the
shape
of
the
dose­
response
curve
in
the
region
of
interest.

Thornton­
Manning
et
al.
(
1994)
administered
bromodichloromethane
to
female
F344
rats
by
aqueous
gavage
for
five
consecutive
days
at
dose
levels
ranging
from
75
to
300
mg/
kg­
day.
This
study
identified
a
NOAEL
of
75
mg/
kg­
day
and
a
LOAEL
of
150
mg/
kg­
day
based
on
increased
liver
and
kidney
weights
and
histologic
lesions
in
the
liver
(
mild
centrilobular
hepatocellular
vacuolar
degeneration)
and
in
the
kidney
(
mild
renal
tubule
vacuolar
degeneration).
An
analogous
study
(
Thornton­
Manning
et
al.,
1994)
conducted
in
female
C57BL/
6J
mice
indicated
that
the
mice
were
less
sensitive
to
bromodichloromethane
than
the
rats,
as
no
treatment­
related
histologic
lesions
were
observed
in
the
liver
or
kidney.
However,
similar
NOAEL
and
LOAEL
values
were
identified
based
on
increased
liver
weight
and
changes
in
serum
chemistry
parameters.
Data
for
renal
tubular
degeneration
in
rats
were
analyzed
using
the
BMD
approach.
BMD
and
BMDL
10
values
of
133
and
65
mg/
kg­
day,
respectively,
were
calculated
for
this
endpoint.
The
BMD
is
in
close
agreement
with
the
BMD
value
calculated
for
the
same
endpoint
using
the
data
of
Lilly
et
al.
(
1994).

Two
reproductive
studies
which
examined
full
litter
resorption
were
also
considered
for
derivation
of
the
One­
day
HA.
Bielmeier
et
al.
(
2001)
examined
the
occurrence
of
full
litter
resorption
in
F344
rats
treated
with
0,
75
or
100
mg/
kg­
day
bromodichloromethane
by
aqueous
gavage
on
gestation
day
9.
The
LOAEL
for
this
effect
was
75
mg/
kg­
day.
Narotsky
et
al.
(
1997)
evaluated
the
same
endpoint
in
F344
rats
administered
0,
25,
50,
or
75
mg/
kg­
day
on
gestation
days
6
through
15.
Full
litter
resorption
was
observed
at
50
and
75
mg/
kg­
day.
The
NOAEL
and
LOAEL
in
this
study
were
thus
identified
as
25
and
50
mg/
kg­
day,
respectively.
When
data
from
these
studies
were
analyzed
using
the
BMD
approach,
BMD
values
of
48
and
23
mg/
kg­
day
were
obtained
for
the
Narotsky
et
al.
(
1997)
and
Bielmeier
et
al.
(
2001)
studies,
respectively.
The
higher
value
from
the
Narotsky
et
al.
(
1997)
study
was
considered
the
more
reliable
estimate
of
the
BMD
because
it
was
based
on
response
data
that
included
lower
doses,
one
of
which
was
an
apparent
NOAEL.
The
BMDL
10
calculated
from
the
Narotsky
et
al.
(
1997)
data
was
30
mg/
kgday

Three
additional
studies
were
considered
as
candidates
for
derivation
of
the
One­
day
HA.
Lilly
et
al.
(
1997)
administered
single
doses
of
bromodichloromethane
by
aqueous
gavage
to
male
F344
rats
at
dose
levels
ranging
from
123
to
492
mg/
kg.
Based
on
changes
in
urinary
parameters,
this
study
identified
a
NOAEL
of
164
mg/
kg­
day
and
a
LOAEL
of
246
mg/
kg­
day.
No
histopathological
examination
was
conducted
in
this
study.
The
study
by
French
et
al.
(
1999),
which
investigated
immune
system
response,
identified
a
similar
NOAEL
value.
However,
the
database
for
immune
response
to
bromodichloromethane
is
limited
when
compared
to
information
on
hepatic
and
renal
toxicity.
Adverse
effects
were
noted
only
at
the
highest
dose
and
frank
effect
level,
and
evidence
for
vehicle
effects
on
immunotoxicity
endpoints
was
observed.
Keegan
et
al.
(
1998)
administered
single
doses
of
bromodichloromethane
to
male
F344
rats
by
gavage
at
dose
levels
ranging
from
21
to
246
mg/
kg.
The
study
authors
identified
a
NOAEL
of
41.0
mg/
kg­
day
and
a
LOAEL
of
81.9
mg/
kg­
day
based
on
elevations
in
serum
markers
of
hepatotoxicity
(
ALT,
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
3
AST,
and
SDH).
Histopathological
examination
was
not
conducted
and
this
was
considered
to
be
a
limitation
of
the
investigation.
These
three
studies
were
not
considered
further
for
derivation
of
the
One­
day
Health
Advisory
(
HA),
and
thus
data
reported
in
them
were
not
analyzed
using
the
BMD
approach.

The
study
conducted
by
Narotsky
et
al.
(
1997)
was
selected
for
derivation
of
the
One­
day
HA.
The
critical
effect
in
this
study
was
full
litter
resorption
observed
in
pregnant
F344
rats
treated
with
bromodichloromethane.
The
BMDL
10
value
calculated
for
this
endpoint
was
30
mg/
kg­
day,
which
is
roughly
half
of
the
most
reliable
BMDL
10
value
calculated
for
histopathological
changes
in
kidney
(
Thornton­
Manning
et
al.,
1994).
Although
dosing
in
the
Narotsky
et
al.
(
1997)
study
lasted
from
gestation
days
6
through
15,
a
subsequent
study
by
Bielmeier
et
al.
(
2001)
indicated
that
a
single
dose
(
75
mg/
kg)
of
bromodichloromethane
on
gestation
day
9
was
sufficient
to
elicit
full
litter
resorption
in
the
same
strain
of
rats.
Since
there
is
presently
insufficient
information
available
to
fully
assess
the
occurrence
of
reproductive
effects
in
humans
exposed
to
bromodichloromethane,
use
of
data
for
full
litter
resorption
was
adopted
as
a
conservative
approach
to
derivation
of
the
One­
day
HA.
The
One­
day
HA
for
a
10­
kg
child
is
calculated
using
the
following
equation:

One­
day
HA
=
(
30
mg/
kg­
day)
(
10
kg)
=
1.0
mg/
L
(
300)
(
1
L/
day)

where:

30
mg/
kg­
day
=
BMDL
10
based
on
incidence
of
full
litter
resorption
in
F344
rats
treated
with
bromodichloromethane
on
gestation
days
6
to
15.

10
kg
=
Assumed
body
weight
of
a
child
300
=
Uncertainty
factor
based
on
NAS/
OW
guidelines.
This
value
includes
a
factor
of
10
to
protect
sensitive
human
populations;
a
factor
of
10
for
extrapolation
from
animals
to
humans;
and
a
factor
of
3
to
account
for
database
limitations
and
uncertainty
regarding
possible
reproductive
effects
of
bromodichloro­
methane
in
humans
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
For
comparative
purposes,
the
One­
day
HA
derived
using
the
conventional
NOAEL/
LOAEL
approach
would
also
be
based
on
data
from
the
Narotsky
et
al.
(
1997)
study.
This
study
identified
a
NOAEL
of
25
mg/
kg­
day
based
on
FLR,
which
was
the
lowest
value
among
the
candidate
studies.
Using
this
NOAEL
and
an
uncertainty
factor
of
300
as
described
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
4
Table
VIII­
1
Summary
of
Candidate
Studies
for
Derivation
of
the
One­
day
HA
for
Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
mg/
kg­
day
LOAEL
mg/
kg­
day
BMD
mg/
kg­
day
BMDL
10
mg/
kg­
day
Lilly
et
al.

(
1994)
Rat
F344
M
6
0
200
400
Gavage
(
oil)
Single
Dose
Body,
liver,
and
kidney
weights,

serum
and
urine
chemistry,
liver
and
kidney
histology
­­
200
(
minimal
renal
tubule
degeneration
and
necrosis,

changes
in
urinary
parameters)
Not
modeled
­­

Lilly
et
al.

(
1994)
Rat
F344
M
6
0
200
400
Gavage
(
aqueous)
Single
Dose
Body,
liver,
and
kidney
weights,

serum
and
urine
chemistry,
liver
and
kidney
histology
­­
200
(
minimal
renal
tubule
degeneration,

changes
in
urinary
parameters)
263
182
(
Hepatic
vacuolar
degeneration
in
males)

131
8.9
(
Renal
tubule
degeneration
in
males)

Lilly
et
al.

(
1997)
Rat
F344
M
5
0
123
164
246
328
492
Gavage
(
aqueous)
Single
Dose
Body,
liver,
and
kidney
weights,

serum
and
urine
chemistry
164
246
(
changes
in
urinary
parameters)
Not
modeled
­­

Keegan
et
al.

(
1998)
Rat
F344
M
6
0
21
31
41
82
123
164
246
Gavage
(
aqueous)
Single
Dose
Body,
liver,
and
kidney
weights,

serum
chemistry
41
82
(
elevated
ALT,

AST,
and
SDH
activities)
Not
modeled
­­
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
mg/
kg­
day
LOAEL
mg/
kg­
day
BMD
mg/
kg­
day
BMDL
10
mg/
kg­
day
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
5
French
et
al.

(
1999)
Rat
C57BL/
6
F
6
0
75
150
300
Gavage
(
aqueous)
5
days
Body,
spleen,
and
thymus
weights,

immune
function
150
300
(
FEL)

(
mortality,

decreased
body
weight,
altered
immune
response)
Not
modeled
­­

Thornton­

Manning
et
al.
(
1994)
Rat
F344
F
6
0
75
150
300
Gavage
(
aqueous)
5
days
Body,
liver,
and
kidney
weights,

serum
chemistry,

liver
and
kidney
histology
75
150
(
increased
liver
and
kidney
weights,
mild
centrilobular
hepatocellular
vacuolar
degeneration,

mild
renal
tubule
vacuolar
degeneration)
133
65
(
renal
tubular
degeneration)

Thornton­

Manning
et
al.
(
1994)
Mouse
C57BL/

6J
F
6
0
75
100
Gavage
(
aqueous)
5
days
Body,
liver,
and
kidney
weights,

serum
chemistry,

liver
and
kidney
histology
75
150
(
increased
liver
weight,
elevated
ALT
and
SDH
activities)
Not
modeled
 
Narotsky
et
al.

(
1997)
*
Rat
F344
F
12­

14
0
25
50
75
Gavage
(
oil)

(
water)**
Gestation
days
6­
15
Body
weight,

clinical
signs,
developmental
parameters
25
50
(
full­
litter
resorption)
48
30
(
full­
litter
resorption)

Bielmeier
et
al.
(
2001)*
Rat
F344
F
8­
11
0
75
100
Gavage
(
aq)
Gestation
day
9
Full
litter
resorption;

hormone
profiles
­­
75
(
full­
litter
resorption)
23
4.2
(
full­
litter
resorption)

*
The
NOAEL
and
LOAEL
values
listed
are
for
reproductive
or
developmental
effects.

**
The
NOAEL
and
LOAEL
values
were
the
same
in
either
vehicle.
BMD
modeling
was
performed
on
aqueous
vehicle
data
only.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
6
above,
the
One­
day
HA
for
a
10­
kg
child
calculated
using
the
conventional
approach
would
be
0.8
mg/
L
(
rounded
from
0.83
mg/
L).

b.
Ten­
day
Health
Advisory
Sixteen
studies
were
considered
for
derivation
of
the
Ten­
day
HA
for
bromodichloromethane.
These
studies
are
summarized
in
Table
VIII­
2
below.
Aida
et
al.
(
1992a)
administered
microencapsulated
bromodichloromethane
in
the
diet
to
Wistar
rats
for
one
month
at
dose
levels
ranging
from
20.6
to
203.8
mg/
kg­
day.
This
study
identified
a
NOAEL
of
61.7
mg/
kg­
day
and
a
LOAEL
of
189.0
mg/
kg­
day
in
male
rats
based
on
histologic
changes
in
the
liver
(
swelling
of
hepatocytes,
single
cell
necrosis,
hepatic
cord
irregularity,
and
bile
duct
proliferation).
Analysis
using
the
BMD
approach
calculated
BMD
and
BMDL
10
values
of
34
and
17
mg/
kg­
day,
respectively,
based
on
data
for
liver
cell
vacuolization
in
females.

Data
from
four
of
the
other
candidate
studies
are
consistent
with
the
histopathological
results
obtained
by
Aida
et
al.
(
1992a).
Melnick
et
al.
(
1998)
administered
bromodichloromethane
by
gavage
to
female
B6C3F
1
mice
for
5
days/
week
for
3
weeks
and
identified
a
NOAEL
of
75
mg/
kg­
day
(
duration­
adjusted
NOAEL
of
54
mg/
kg­
day)
and
a
LOAEL
of
150
mg/
kg­
day
(
duration­
adjusted
LOAEL
of
107
mg/
kg­
day)
based
on
histologic
changes
in
the
liver
(
hepatocyte
hydropic
degeneration).
Analysis
using
the
BMD
approach
calculated
duration­
adjusted
BMD
and
BMDL
10
values
of
31
and
8.4
mg/
kg­
day,
respectively,
for
this
endpoint.
Condie
et
al.
(
1983)
administered
bromodichloromethane
by
gavage
to
male
CD­
1
mice
for
14
days
and
identified
a
NOAEL
of
74
mg/
kg­
day
and
a
LOAEL
of
148
mg/
kg­
day
based
on
minimal
to
moderate
liver
and
kidney
lesions.
Analysis
using
the
BMD
approach
calculated
BMD
and
BMDL
10
values
of
24
and
7.5
mg/
kg­
day,
respectively,
based
on
data
for
histopathological
changes
in
the
liver.
NTP
(
1998)
conducted
histopathologic
examinations
in
conjunction
with
a
study
of
reproductive
and
developmental
toxicity
in
Sprague­
Dawley
rats.
Although
no
reproductive
or
developmental
toxicity
was
observed
at
the
dose
levels
investigated,
histopathological
changes
were
noted
in
the
liver
of
males
rats
treated
with
the
compound
for
35
days.
The
NOAEL
and
LOAEL
for
this
effect
were
9
and
38
mg/
kg­
day,
respectively.
Analysis
of
data
for
single
cell
hepatic
necrosis
using
the
BMD
approach
calculated
BMD
and
BMDL
10
values
of
35
and
18
mg/
kg­
day,
respectively,
which
are
virtually
identical
to
the
values
calculated
using
the
liver
cell
vacuolization
data
for
females
from
the
Aida
et
al.
(
1992b)
study.
Coffin
et
al.
(
2000)
observed
hydropic
degeneration
in
female
mice
treated
with
150
mg/
kg­
day
bromodichloromethane
in
corn
oil
for
11
days.
These
data
were
not
modeled
because
other
studies
utilized
doses
lower
than
150
mg/
kg­
day,
which
allowed
better
characterization
of
response
in
the
low
dose
region
of
the
dose­
response
curve.

In
contrast
to
the
studies
described
above,
Chu
et
al.
(
1982a)
did
not
observe
any
microscopic
lesions
in
the
liver
when
Sprague­
Dawley
rats
were
administered
bromodichloromethane
at
doses
up
to
68
mg/
kg­
day
in
the
drinking
water
for
28
days.
The
studies
of
Munson
et
al.
(
1982)
and
NTP
(
1987)
did
not
conduct
histopathological
examinations.
These
studies
identified
NOAELs
ranging
from
50
to
150
mg/
kg­
day
and
LOAELs
ranging
from
125
to
300
mg/
kg­
day
for
other
endpoints,
including
depressed
humoral
immunity
(
Munson
et
al.,
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
7
1982),
decreased
weight
gain
(
NTP,
1987;
rats),
and
increased
mortality
and
gross
renal
pathology
(
NTP,
1987;
mice).
These
data
were
not
analyzed
by
the
BMD
approach.

Seven
studies
that
examined
developmental
and/
or
reproductive
endpoints
were
evaluated.
Two
studies
reported
the
incidence
of
full
litter
resorption
in
F344
rats
following
treatment
with
bromodichloromethane.
Bielmeier
et
al.
(
2001)
examined
the
occurrence
of
full
litter
resorption
in
F344
rats
treated
with
0,
75
or
100
mg/
kg­
day
bromodichloromethane
by
aqueous
gavage
on
gestation
day
9.
The
LOAEL
for
this
effect
was
75
mg/
kg­
day.
Narotsky
et
al.
(
1997)
evaluated
the
same
endpoint
in
F344
rats
administered
0,
25,
50,
or
75
mg/
kg­
day
on
gestation
days
6
through
15.
Full
litter
resorption
was
observed
at
50
and
75
mg/
kg­
day.
The
NOAEL
and
LOAEL
in
this
study
were
thus
identified
as
25
and
50
mg/
kg­
day,
respectively.
When
data
from
these
studies
were
analyzed
using
the
BMD
approach,
BMD
values
of
48
and
23
mg/
kg­
day
were
obtained
for
the
Narotsky
et
al.
(
1997)
and
Bielmeier
et
al.
(
2001)
studies,
respectively.
The
higher
value
from
the
Narotsky
et
al.
(
1997)
study
was
considered
the
more
reliable
estimate
of
the
BMD
because
it
was
based
on
response
data
that
included
lower
doses,
one
of
which
was
an
apparent
NOAEL.
The
BMDL
10
calculated
from
the
Narotsky
et
al.
(
1997)
data
was
30
mg/
kgday
Ruddick
et
al.
(
1983)
observed
an
increased
incidence
of
sternebral
aberrations
in
the
pups
of
Sprague­
Dawley
rats
administered
bromodichloromethane
in
corn
oil
by
gavage.
Statistical
analysis
of
the
published
data
indicated
that
the
NOAEL
and
LOAEL
for
this
effect
were
100
mg/
kg­
day
and
200
mg/
kg­
day,
respectively.
The
BMD
and
BMDL
10
obtained
for
this
study
were
27
and
15
mg/
kg­
day,
respectively.

The
remaining
reproductive/
developmental
studies
sponsored
by
the
Chlorine
Chemistry
Council
(
CCC)
were
also
evaluated.
CCC
(
2000a,
b)
examined
developmental
toxicity
in
New
Zealand
rabbits
and
identified
developmental
NOAELs
of
76
and
55
mg/
kg­
day
(
the
highest
doses
tested
in
each
study).
The
CCC
(
CCC,
2000c,
d)
also
examined
reproductive
and
developmental
toxicity
in
Sprague
Dawley
rats.
In
a
range­
finding
study
(
CCC,
2000c),
F
1
generation
pups
exposed
to
bromodichloromethane
via
lactation
and
possibly
by
consumption
of
water
supplied
to
the
dams
exhibited
reduced
body
weights
and
body
weight
gains.
These
effects
occurred
at
exposure
levels
which
also
resulted
in
maternal
toxicity.
Biologically
meaningful
average
daily
doses
could
not
be
established
in
this
experiment;
therefore,
the
concentration­
based
NOAEL
and
LOAEL
values
for
developmental
effects
were
50
and
150
ppm.
based
on
changes
in
F
1
pup
body
weight
and
body
weight
gain.
In
a
subsequent
developmental
study,
CCC
(
2000d)
identified
NOAEL
and
LOAEL
values
of
45
and
82
mg/
kg­
day,
respectively,
based
on
decreased
number
of
ossification
sites
per
fetus
for
the
forelimb
phalanges
and
hindlimb
metatarsals
and
phalanges.
This
effect
was
observed
at
doses
associated
with
maternal
toxicity.
Endpoints
from
these
studies
were
not
modeled
because
other
studies
identified
adverse
effects
at
lower
doses.

Data
for
maternal
toxicity
from
three
reproductive/
developmental
studies
were
also
considered
for
derivation
of
the
10­
day
HA
for
bromodichloromethane.
Narotsky
et
al.
(
1997)
observed
decreased
maternal
body
weight
gain
on
gestation
days
6
to
8
in
female
rats
administered
25
mg/
kg­
day
(
the
lowest
dose
tested)
by
aqueous
gavage
in
10%
Emulphor.
BMD
modeling
identified
BMD
and
BMDL
10
values
of
18
and
10
mg/
kg­
day,
respectively,
for
this
endpoint.
The
data
reported
in
this
study
did
not
permit
evaluation
of
body
weight
or
body
Draft
­
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Not
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or
Quote
February
20,
2003
VIII
­
8
weight
gain
at
other
time
points
during
the
treatment
period.
CCC
(
2000d)
reported
decreased
maternal
body
weight
gain
at
several
time
points
in
pregnant
rats
administered
bromodichloromethane
in
drinking
water,
with
the
most
severe
effect
observed
immediately
after
initiation
of
treatment
on
gestation
days
6
to
7.
The
NOAEL
and
LOAEL
for
decreased
body
weight
on
gestation
days
6
to
7
were
18.4
and
45
mg/
kg­
day,
respectively.
When
body
weight
gain
for
this
interval
was
modeled,
the
resulting
BMD
and
BMDL
10
values
were
approximately
18
and
15
mg/
kg­
day,
respectively.
However,
the
modeled
fits
to
the
data
were
poor,
and
the
results
were
not
considered
sufficiently
reliable
for
derivation
of
a
health
advisory.
To
address
this
problem,
body
weight
gain
data
for
gestation
days
6
to
9
were
also
modeled.
Reliable
values
of
23
and
18
mg/
kg­
day
were
obtained
for
the
BMD
and
BMDL
10,
respectively.
The
CCC
(
2000b)
study
observed
decreased
maternal
body
weight
gain
at
several
time
points
in
pregnant
rabbits
administered
bromodichloromethane
in
the
drinking
water
on
gestation
days
6
to
29.
The
NOAEL
and
LOAEL
for
decreased
maternal
body
weight
gain
(
corrected
for
gravid
uterine
weight)
on
gestation
days
6
to
21
were
13.4
and
35.3
mg/
kg­
day,
respectively.
A
BMD
value
of
50
mg/
kg­
day
was
obtained
for
this
data
set,
but
the
BMDS
software
failed
to
identify
the
corresponding
BMDL
10.
No
further
modeling
was
attempted
since
this
value
was
well
above
the
lowest
BMDs
obtained
in
some
other
candidate
studies.

As
evident
from
the
data
in
Table
VIII­
2,
the
four
studies
that
examined
histopathological
changes
in
the
liver
are
in
close
agreement,
having
identified
BMD
values
ranging
from
24
to
35
mg/
kg­
day.
The
corresponding
BMDL
10
values
ranged
from
7.5
to
18
mg/
kg­
day.
Maternal
toxicity
occurred
in
rats
at
similar
levels
in
two
developmental
studies.
These
studies
identified
BMD
values
of
18
and
23
mg/
kg­
day,
with
corresponding
BMDL
10
values
of
10
and
18
mg/
kgday
The
NTP
(
1998)
and
CCC
(
2000d)
drinking
water
studies
were
selected
to
derive
the
Tenday
HA.
Selection
of
these
studies
was
based
on
the
administration
of
bromodichloromethane
in
drinking
water,
the
most
relevant
route
of
exposure.
In
addition,
these
studies
utilized
a
lower
range
of
doses,
which
provided
information
on
the
shape
of
the
dose­
response
curve
in
the
region
of
interest.
The
Ten­
day
HA
is
calculated
according
to
the
following
equation:

Ten­
day
HA
=
(
18
mg/
kg­
day)
(
10
kg)
=
0.60
mg/
L
(
rounded
to
0.6
mg/
L)
(
300)
(
1
L/
day)

where:

18
mg/
kg­
day
=
BMDL
10
based
on
single
cell
hepatic
necrosis
in
rats
administered
bromodichloromethane
in
the
drinking
water
for
35
days
or
decreased
maternal
body
weight
gain
on
gestation
days
6­
9
in
pregnant
female
rats
administered
bromodichloromethane
in
the
drinking
water.

10
kg
=
Assumed
body
weight
of
a
child
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
9
300
=
Uncertainty
factor
based
on
NAS/
OW
guidelines.
This
value
includes
a
factor
of
10
to
protect
sensitive
human
populations
and
a
factor
of
10
for
extrapolation
from
animals
to
humans,
and
a
factor
of
3
to
account
for
database
limitations
and
uncertainty
regarding
possible
reproductive
effects
of
bromodichloromethane
in
humans
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
For
comparative
purposes,
the
Ten­
day
HA
derived
using
the
conventional
NOAEL/
LOAEL
approach
would
be
based
on
data
from
the
CCC
(
2000d)
study.
This
study
identified
a
NOAEL
of
18
mg/
kg­
day
based
on
reduced
maternal
body
weight
gain
in
pregnant
rabbits.
Using
this
NOAEL
and
an
uncertainty
factor
of
300
as
described
above,
the
Ten­
day
HA
for
a
10­
kg
child
calculated
using
the
conventional
approach
would
be
0.60
mg/
L
(
rounded
to
0.6
mg/
L).
Draft
­
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Not
Cite
or
Quote
February
20,
2003
VIII
­
10
Table
VIII­
2
Summary
of
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Aida
et
al.

(
1992a)
Rat
Wistar
M,
F
7
Male
0
21
62
189
Females
0
21
66
204
Feed
1
month
Clinical
signs,
body
weight,
serum
chemistry,
hematology,
histology
62
189
(
liver
histopathology
in
males)
34
17
(
liver
cell
vacuolation
in
females)

Chu
et
al.

(
1982a)
Rat
SD
M
10
0
0.8
8
68
Drinking
water
28
days
Clinical
signs,
serum
chemistry,
histology
68
­­
No
data
to
model
­­

Condie
et
al.

(
1983)
Mouse
CD­
1
M
8­
16
0
37
74
148
Gavage
(
oil)
14
days
Serum
enzymes,
PAH
uptake
in
vitro,

histology
74
148
(
elevated
ALT,

decreased
PAH
uptake,
liver
and
kidney
histopathology)
24
7.5
(
hepatic
centrilobular
pallor)

125
53
(
Renal
epithelial
hyperplasia)

Melnick
et
al.

(
1998)
Mouse
B6C3F
1
F
10
0
75
150
326
Gavage
(
oil)
3
weeks
(
5
d/
wk)
Body
and
liver
weights,
serum
chemistry,
liver
histology
75
150
(
liver
histopathology)
31
*
8.4
(
Hepatocyte
hydropic
degeneration)

Munson
et
al.

(
1982)
Mouse
CD­
1
M,
F
8­
12
Males
0
50
125
250
Gavage
(
aq.)
14
days
Body
and
organ
weights,
serum
chemistry,
hematology,
and
immune
function
50
125
(
depressed
humoral
immunity)
Not
modeled
­­
Table
VIII­
2
(
cont.)

Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
11
NTP
(
1987)
Rat
F344/
N
M,
F
5
0
38
75
150
300
600
Gavage
(
oil)
14
days
Body
weight,
clinical
signs,
gross
necropsy
150
300
(
decreased
weight
gain)
Not
modeled
­­

NTP
(
1987)
Mouse
B6C3F
1
M,
F
5
0
19
38
75
150
300
Gavage
(
oil)
14
days
Body
weight,
clinical
signs,
gross
necropsy
75
150
(
FEL)
*

(
mortality,

lethargy,
gross
renal
pathology)
Not
modeled
­­

Coffin
et
al.

(
2000)
Mouse
B6C3F
1
F
10
0
150
300
Gavage
(
oil)
11
days
Relative
liver
wt.,

liver
histopathology;

labeling
index
­­
150
Not
modeled
­­

NTP
(
1998)
Rat
SD
M
(
group
A)
5­
13
0
9
38
67
Drinking
water
35
days
Body
and
organ
weights,
serum
chemistry,
hematology,
gross
necropsy,
histology,

sperm
evaluation
9
38
(
liver
histopathology)
35
18
(
liver
cell
necrosis)

Ruddick
et
al.
(
1983)*
Rat
SD
F
9­
14
0
50
100
200
Gavage
(
oil)
Gestation
days
6­
15
Body
and
organ
weights;
maternal
serum
chemistry;

hematology,
and
histopathology;

developmental
parameters
100
200
(
sternebral
aberrations)
27
15
(
sternebral
aberrations)
Table
VIII­
2
(
cont.)

Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
12
Narotsky
et
al.
(
1997)*
Rat
F344
F
12­

14
0
25
50
75
Gavage
(
oil)

(
Emulphor
Gestation
days
6­
15
Body
weight,
clinical
signs,
developmental
parameters
25
(
developmental
)
50
(
full­
litter
resorption)
48
30
(
full­
litter
resorption)

­

(
maternal)
25
(
reduced
matermal
body
weight
gain
gestation
days
6­
8,
aqueous
vehicle
only)
18
10
(
reduced
matermal
body
weight
gain
gestation
days
6­

8,
aqueous
vehicle
only)

Bielmeier
et
al.
(
2001)*
Rat
F344
F
8­
11
0
75
100
Gavage
(
aq)
Gestation
day
9
Full
litter
resorption;

hormone
profiles
­­
75
(
full­
litter
resorption)
23
4.2
(
full­
litter
resorption)

CCC
(
2000c)*
Rat
SD
M,
F
10
0
ppm
50
ppm
150
ppm
450
ppm
1350
ppm
Drinking
water
Males
64
days
Females
74
days
Reproductive
and
developmental
parameters
50
ppm
150
ppm
(
reduced
F
1
pup
weight
and
weight
gain)
Not
modeled
­­

CCC
(
2000d)*
Rat
SD
F
25
0.0
2.2
18.4
45.0
82.0
Drinking
water
Gestation
days
6­
21
Reproductive
and
developmental
parameters
45.3
(
developmental)
82.0
(
reduced
number
of
ossification
sites
in
phalanges
or
metatarsals
occurring
with
maternal
toxicity)
Not
modeled
­­

18.4
(
maternal)
45
(
reduced
maternal
body
weight
gain
gestation
days
6­
7)
23
18
(
reduced
maternal
body
weight
gain
gestation
days
6­

9;
see
text
for
comment)
Table
VIII­
2
(
cont.)

Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
13
CCC
(
2000a)*
Rabbit
NZW
F
5
0.0
13.9
32.3
76.3
Drinking
water
Gestation
days
6­
29
Reproductive
and
developmental
parameters
76.3
(
developmental)
­
Not
modeled
­­

CCC
(
2000b)*
Rabbit
NZW
F
25
0
1.4
13.4
35.6
55.3
Drinking
Water
Gestation
days
6­
29
Clinical
sign,
gross
lesions,
reproductive
and
developmental
endpoints
55
(
developmental)
­
Not
modeled
­­

13.4
(
maternal)
35.3
(
reduced
corrected
maternal
body
weight
gain
gestation
days
6­
29)
50
BMDS
software
failed
*
The
NOAEL
and
LOAEL
values
listed
are
for
reproductive
or
developmental
effects.

**
The
NOAEL
and
LOAEL
values
were
the
same
for
developmental
effects
in
either
vehicle.
The
LOAEL
for
maternal
toxicity
was
25
mg/
kg­
day
for
the
aqueous
vehicle
(
10%
Emulphor).
The
NOAEL
and
LOAEL
for
maternal
toxicity
using
the
corn
oil
vehicle
were
25
mg/
kg­
day
and
50
mg/
kg­
day,

respectively.
BMD
modeling
was
performed
on
aqueous
vehicle
data
only.

*
BMD
and
BMDL10
calculated
using
duration
adjusted
doses
Abbreviations:
FEL,
Frank
effect
level;
SD,
Sprague­
Dawley;
NZW,
New
Zealand
White
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
14
c.
Longer­
term
Health
Advisory
Two
rodent
oral
exposure
studies
conducted
by
NTP
(
1987)
were
considered
for
derivation
of
the
Longer­
term
HA
for
bromodichloromethane.
In
addition,
eight
reproductive
studies
were
considered.
These
studies
are
summarized
in
Table
VIII­
3
below.

NTP
(
1987)
administered
bromodichloromethane
by
gavage
to
F344/
N
rats
for
5
days/
week
for
13
weeks
at
dose
levels
ranging
from
19
to
300
mg/
kg­
day.
Based
on
decreased
weight
gain,
this
study
identified
a
NOAEL
of
75
mg/
kg­
day
and
a
LOAEL
of
150
mg/
kg­
day.
Treatment­
related
hepatic
and
renal
lesions
were
observed
only
at
the
high
dose.
In
a
similar
study,
NTP
(
1987)
administered
bromodichloromethane
by
gavage
to
B6C3F
1
mice
for
5
days/
week
for
13
weeks
at
dose
levels
ranging
from
6.25
to
100
mg/
kg­
day
for
males
and
from
25
to
400
mg/
kg­
day
for
females.
This
study
identified
a
NOAEL
of
50
mg/
kg­
day
and
a
LOAEL
of
100
mg/
kg­
day
based
on
histologic
alterations
in
the
kidney
(
focal
necrosis
of
the
proximal
renal
tubular
epithelium
and
nephrosis)
of
male
mice.
BMD
analysis
using
the
BMDS
program
identified
duration­
adjusted
BMD
values
of
63
and
75
mg/
kg­
day
for
focal
necrosis
of
renal
tubular
epithelium
in
males
and
vacuolated
cytoplasm
in
the
liver
of
females,
respectively.
The
corresponding
BMDL
10
values
for
these
renal
and
hepatic
effects
were
35
and
47
mg/
kg­
day,
respectively.

Eight
reproductive
or
developmental
studies
(
Ruddick
et
al.,
1983;
Narotsky
et
al.,
1997;
Bielmeier
et
al.
2001;
CCC,
2000a,
b,
c,
d;
CCC,
2002)
were
also
considered
for
derivation
of
the
Longer­
term
HA.
CCC
(
2002)
identified
a
LOAEL
of
150
ppm
(
approximately
11.6
to
40.2
mg/
kg­
day)
for
delayed
sexual
maturation
in
F
1
male
rats
in
a
two­
generation
study
of
bromodichloromethane
administered
in
drinking
water.
The
LOAEL
for
parental
effects
in
the
study
was
also
150
ppm,
based
on
decreased
body
weight
and
body
weight
gain
in
F
1
females
and
F
1
males
and
females.
Bielmeier
et
al.
(
2001)
examined
the
occurrence
of
full
litter
resorption
in
F344
rats
treated
with
0,
75
or
100
mg/
kg­
day
bromodichloromethane
by
aqueous
gavage
on
gestation
day
9.
The
LOAEL
for
this
effect
was
75
mg/
kg­
day.
Narotsky
et
al.
(
1997)
evaluated
the
same
endpoint
in
F344
rats
administered
0,
25,
50,
or
75
mg/
kg­
day
on
gestation
days
6
through
15.
Full
litter
resorption
was
observed
at
50
and
75
mg/
kg­
day.
The
NOAEL
and
LOAEL
in
this
study
were
thus
identified
as
25
and
50
mg/
kg­
day,
respectively.
When
data
from
these
studies
were
analyzed
using
the
BMD
approach,
BMD
values
of
48
and
23
mg/
kg­
day
were
obtained
for
the
Narotsky
et
al.
(
1997)
and
Bielmeier
et
al.
(
2001)
studies,
respectively.
The
higher
value
from
the
Narotsky
et
al.
(
1997)
study
was
considered
the
more
reliable
estimate
of
the
BMD
because
it
was
based
on
response
data
that
included
lower
doses,
one
of
which
was
an
apparent
NOAEL.
The
BMDL
10
calculated
from
the
Narotsky
et
al.
(
1997)
data
was
30
mg/
kgday
Studies
conducted
by
NTP
(
1998)
did
not
detect
reproductive
or
developmental
toxicity
at
doses
up
to
116
mg/
kg­
day.
Two
studies
conducted
in
New
Zealand
White
rabbits
did
not
detect
developmental
effects
at
doses
up
to
55
and
76
mg/
kg­
day,
respectively
(
CCC,
2000a,
b).

Three
additional
studies
in
rats
identified
developmental
effects
that
occurred
at
dose
levels
that
also
resulted
in
maternal
toxicity.
In
a
range­
finding
study
(
CCC,
2000c),
F
1
generation
pups
exposed
to
bromodichloromethane
via
lactation
and
possibly
by
consumption
of
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
15
water
supplied
to
the
dams
exhibited
reduced
body
weights
and
body
weight
gains.
Biologically
meaningful
daily
doses
could
not
be
established
in
this
experiment;
therefore,
the
concentrationbased
NOAEL
and
LOAEL
values
are
50
ppm
and
150
ppm.
based
on
reduced
body
weight
and
body
weight
gain
in
the
F
1
pups.
In
a
subsequent
developmental
study,
CCC
(
2000d)
identified
NOAEL
and
LOAEL
values
of
45
and
82
mg/
kg­
day,
respectively,
based
on
decreased
number
of
ossification
sites
per
fetus
for
the
forelimb
phalanges
and
hindlimb
metatarsals
and
phalanges.
These
reversible
developmental
delays
occurred
at
doses
which
also
resulted
in
maternal.
Endpoints
from
these
studies
were
not
modeled
because
other
effects
were
observed
at
lower
doses.
Ruddick
et
al.
(
1983)
observed
an
increased
incidence
of
sternebral
aberrations
in
the
pups
of
Sprague­
Dawley
rats
administered
bromodichloromethane
in
corn
oil
by
gavage.
Statistical
analysis
of
the
published
data
indicated
that
the
NOAEL
and
LOAEL
for
this
effect
were
100
mg/
kg­
day
and
200
mg/
kg­
day,
respectively.
The
lowest
dose
tested
was
50
mg/
kg­
day.
The
BMD
and
BMDL
10
obtained
for
this
study
were
27
and
15
mg/
kg­
day,
respectively.
However,
examination
of
the
modeling
output
indicated
that
none
of
the
available
models
fit
the
data
well
in
the
low­
dose
region
of
the
curve.
Therefore,
the
reliability
of
these
values
is
questionable.

The
Narotsky
et
al.
(
1997)
and
CCC
(
2000d)
studies
identified
BMDL
10
values
of
10
and
18
mg/
kg­
day
based
on
reduced
maternal
body
weight
gain.
The
CCC
data
were
considered
the
most
relevant
since
they
were
obtained
from
a
drinking
water
study
which
utilized
concentrations
that
resulted
in
daily
doses
well
below
those
used
in
the
Narotsky
study.
The
NTP
(
1987)
and
the
Narotsky
et
al.
(
1997)
studies
provided
similar,
but
higher,
BMDL
10
values,
based
on
reproductive
and
histopathological
endpoints.
The
NTP
(
1987)
study
utilized
bromodichloromethane
doses
as
low
as
6.3
mg/
kg­
day,
in
contrast
to
the
reproductive
study
conducted
by
Narotsky
et
al.
(
1997)
in
which
the
lowest
dose
was
25
mg/
kg­
day.
The
NTP
(
1987)
data
thus
provide
more
information
about
the
shape
of
the
dose­
response
curve
in
the
region
of
interest.
The
BMD
data
for
focal
necrosis
of
renal
tubular
epithelium
and
reduced
body
weight
gain
in
pregnant
female
rats
were
thus
selected
as
the
most
reliable
basis
for
determining
the
Longer­
term
HA.
Using
the
lower
of
the
two
values,
the
duration­
adjusted
BMDL
10
of
18
mg/
kg­
day
for
reduced
maternal
body
weight
gain,
the
Longer­
term
HA
for
a
10
kg
child
is
calculated
according
to
the
following
equation:

Longer­
term
HA
=
(
18
mg/
kg­
day)
(
10
kg)
=
0.60
mg/
L
(
rounded
to
0.6
mg/
L)
(
300)
(
1
L/
day)

where:

18
mg/
kg­
day
=
BMDL
10
based
on
decreased
maternal
body
weight
gain
on
gestation
days
6­
9
in
pregnant
female
rats
administered
bromodichloromethane
in
the
drinking
water.

10
kg
=
Assumed
body
weight
of
a
child
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
16
300
=
Uncertainty
factor
based
on
NAS/
OW
guidelines.
This
value
includes
a
factor
of
10
to
protect
sensitive
human
populations
and
a
factor
of
10
for
extrapolation
from
animals
to
humans,
and
a
factor
of
3
to
account
for
uncertainty
regarding
possible
reproductive
effects
of
bromodichloromethane
in
humans
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
The
Longer­
term
HA
for
adults
would
be
calculated
as
follows:

Longer­
term
HA
=
(
18
mg/
kg­
day)
(
70
kg)
=
2.1
mg/
L
(
rounded
to
2
mg/
L)
(
300)
(
2
L/
day)

where:

18
mg/
kg­
day
=
BMDL
10
based
on
reduced
body
weight
gain
in
female
rats
administered
bromodichloromethane
in
the
drinking
water
70
kg
=
Assumed
body
weight
of
an
adult
300
=
Uncertainty
factor
based
on
NAS/
OW
guidelines.
This
value
includes
a
factor
of
10
to
protect
sensitive
human
populations
and
a
factor
of
10
for
extrapolation
from
animals
to
humans,
and
a
factor
of
3
to
account
for
database
limitations
and
uncertainty
related
to
potential
reproductive
effects
of
bromodichloromethane
in
humans
2
L/
day
=
Assumed
water
consumption
of
a
70­
kg
adult
For
purposes
of
comparison
,
a
Longer­
term
HA
derived
using
the
conventional
NOAEL/
LOAEL
approach
would
be
based
on
the
study
conducted
by
CCC
(
2000d).
This
study
identified
a
NOAEL
of
18.4
mg/
kg­
day
and
a
LOAEL
of
45
mg/
kg­
day
based
on
reduced
body
weight
gain
on
gestation
days
6
to
7.
Using
the
NOAEL
of
18.4
mg/
kg­
day,
and
assuming
drinking
water
ingestion
of
1
L/
day
and
an
uncertainty
factor
of
300
(
including
factors
of
10
for
interspecies
extrapolation
and
protection
of
susceptible
populations,
and
a
factor
of
3
for
database
limitations
and
uncertainty
regarding
potential
reproductive
effects
in
humans),
the
Longer­
term
HA
for
a
10
kg
child
would
be
0.6
mg/
L.
The
Longer­
term
HA
for
a
70
kg
adult
consuming
2
L/
day
would
be
2
mg/
kg­
day.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
17
Table
VIII­
3
Summary
of
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
for
Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
*

NTP
(
1987)
Rat
F344/
N
M,
F
10
0
19
38
75
150
300
Gavage
(
oil)
13
weeks
(
5
d/
wk)
Body
weight,

clinical
signs,

histology
75
150
(
decreased
weight
gain)

(
hepatic
and
renal
lesions
at
300)
Not
modeled
­­

NTP
(
1987)
Mouse
B6C3F
1
M,
F
10
Male
0
6.3
13
25
50
100
Female
0
25
50
100
200
400
Gavage
(
oil)
13
weeks
(
5
d/
wk)
Body
weight,

clinical
signs,

histology
50
100
(
renal
lesions)
63
35
(
focal
necrosis
of
renal
tubular
epithelium
in
males)

75
47
(
Hepatic
vacuolated
cytoplasm
in
females)

Ruddick
et
al.
(
1983)*
Rat
SD
F
9­
14
0
50
100
200
Gavage
(
oil)
Gestation
days
6­
15
Body
and
organ
weights;
maternal
serum
chemistry;
hematology,
and
histopathology;

developmental
parameters
100
200
(
increased
incidence
of
sternebral
variations)
27
15
(
increased
incidence
of
sternebral
variations)

Narotsky
et
al.

(
1997)
*
Rat
F344
F
12­

14
0
25
50
75
Gavage
(
oil)

(
aq)**
Gestation
days
6­
15
Body
weight,

clinical
signs,
developmental
parameters
25
(
developmental)
50
(
full­
litter
resorption)
48
30
(
full­
litter
resorption)

­

(
maternal)
25
(
reduced
maternal
body
weight
gain
gestation
days
6­

8,
aq.
vehicle)
18
10
(
reduced
maternal
body
weight
gain,

gestation
days
6­

8,
aq.
vehicle)
Table
VIII­
3
(
cont.)

Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
*

Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
18
Bielmeier
et
al.
(
2001)*
Rat
F344
F
8­
11
0
75
100
Gavage
(
aq)
Gestation
day
9
Full
litter
resorption;

hormone
profiles
­­
75
(
full­
litter
resorption)
23
4.2
(
full­
litter
resorption)

CCC
(
2000c)*
Rat
SD
M,
F
10
0
ppm
50
ppm
150
ppm
450
ppm
1350
ppm
Drinking
water
Males
64
days
Females
74
days
Reproductive/
developmental
parameters
Developmental
50
ppm
Parental
50
ppm
Developmental
150
ppm­

Parental
50
ppm
Not
modeled
­­

CCC
(
2000d)*
Rat
SD
F
25
0.0
2.2
18.4
45.0
82.0
Drinking
water
Gestation
days
6­
21
Reproductive/
developmental
parameters
45.3
(
developmental)
82.0
(
reduced
number
of
ossification
sites
in
phalanges
or
metatarsals
occurring
with
maternal
toxicity)
Not
modeled
­­

18.4
(
maternal)
45
(
reduced
maternal
body
weight
gain
gestation
days
6­

7)
23
18
(
reduced
maternal
body
weight
gain
gestation
days
6­

9;
see
comments
in
text)

CCC
(
2002)*
Rat
SD
M,
F
30
0
ppm
50
ppm
150
ppm
450
ppm
Drinking
water
Two
generations
Reproductive
parameters
50
ppm
(
offspring)
150
ppm
(
delayed
sexual
maturation
in
F
1
males)
Not
modeled
 
50
ppm
(
parental)
150
ppm
(
Reduced
body
wt
and
body
wt
gain
in
F
0
females
and
F
1
males
and
females)
Table
VIII­
3
(
cont.)

Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
*

Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
19
CCC
(
2000a)*
Rabbit
NZW
F
5
0
4.9
13.9
32.3
76.3
Drinking
Water
Gestation
day
6
to
29
Body
weight,

clinical
signs,

reproductive
and
developmental
parameters
76.3
(
developmental)
­
Not
modeled
­­

CCC
(
2000b)*
Rabbit
NZW
F
25
0
1.4
13.4
35.6
55.3
Drinking
Water
Gestation
days
6­
29
Clinical
sign,

gross
lesions,

reproductive
and
developmental
endpoints
55.3
(
developmental)
­
Not
modeled
­­

13.4
(
maternal)
35.3
(
reduced
maternal
body
weight
gain
50
(
developmental)
BMD
software
failed
*
BMDL10
value
was
derived
using
duration­
adjusted
doses.

*
Modeled
using
Crump
Benchmark
Dose
Software
*
The
NOAEL
and
LOAEL
values
listed
are
for
reproductive/
developmental
effects.

**
The
developmental
NOAEL
and
LOAEL
values
were
the
same
in
either
vehicle.
The
LOAELs
for
maternal
toxicity
were
25
mg/
kg­
day
and
50
mg/

kgday
for
the
aqueous
and
corn
oil
vehicles
respectively.
BMD
modeling
was
performed
on
aqueous
vehicle
data
only.

Abbreviations:
NA,
Not
available;
SD,
Sprague­
Dawley;
NZW,
New
Zealand
White
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
20
d.
Reference
dose,
Drinking
Water
Equivalent
Level
and
Lifetime
Health
Advisory
This
section
reports
the
existing
RfD
value
for
bromodichloromethane
and
describes
the
derivation
of
the
RfD
for
this
compound.
This
section
also
describes
the
calculation
of
Drinking
Water
Equivalent
Level
and
Lifetime
Health
Advisory
values
which
require
the
RfD
as
input.
For
this
document,
new
and
existing
studies
were
reviewed
and
appropriate
candidate
data
were
selected
for
benchmark
dose
(
BMD)
modeling.
The
results
of
BMD
modeling
were
used
in
conjunction
with
appropriate
uncertainty
factors
to
calculate
the
RfD.
A
comparison
of
the
RfD
derived
using
the
BMD
approach
to
the
results
obtained
using
the
conventional
NOAEL/
LOAEL
approach
is
also
provided.

Description
of
the
Existing
RfD
The
existing
RfD
for
bromodichloromethane
is
0.02
mg/
kg­
day
(
IRIS,
1993a).
This
value
was
derived
using
a
duration­
adjusted
LOAEL
of
17.9
mg/
kg­
day
identified
for
renal
cytomegaly
in
B6C3F
1
mice
administered
bromodichloromethane
by
corn
oil
gavage
for
5
days/
week
for
102
weeks
(
NTP,
1987).
An
uncertainty
factor
of
100
was
used
to
account
for
extrapolation
from
animal
data
and
for
protection
of
sensitive
human
subpopulations.
An
additional
factor
of
10
was
used
because
the
RfD
was
based
on
a
LOAEL
(
although
it
was
considered
minimally
adverse)
and
to
account
for
lack
of
reproductive
data.

Identification
of
Candidate
Studies
for
Derivation
of
the
RfD
Several
studies
of
chronic
duration
were
considered
for
derivation
of
the
RfD
for
bromodichloromethane.
These
studies
are
summarized
in
Table
VIII­
4
below.
NTP
(
1987)
administered
bromodichloromethane
to
F344/
N
rats
by
gavage
in
corn
oil
at
doses
of
50
or
100
mg/
kg­
day
for
5
day/
week
for
102
weeks.
This
study
identified
a
LOAEL
of
50
mg/
kg­
day
based
on
histologic
alterations
in
the
liver
and
kidney.
In
a
similar
study,
NTP
(
1987)
administered
bromodichloromethane
by
gavage
in
corn
oil
to
B6C3F
1
mice
for
5
days/
week
for
102
weeks
at
dose
levels
of
25
or
50
mg/
kg­
day
for
males
and
75
or
150
mg/
kg­
day
for
females.
Based
on
histologic
alterations
in
the
liver,
kidney,
and
thyroid
of
male
mice,
this
study
identified
a
LOAEL
of
25
mg/
kg­
day,
which
is
consistent
with
the
value
identified
in
the
rat
study.
In
a
third
study,
Tobe
et
al.
(
1982)
administered
microencapsulated
bromodichloromethane
to
Wistar
rats
in
the
diet
at
dose
levels
ranging
from
6
to
168
mg/
kg­
day.
Histologic
data
for
the
animals
exposed
to
bromodichloromethane
were
reported
by
Aida
et
al.
(
1992b).
This
study
identified
a
LOAEL
for
male
rats
of
6
mg/
kg­
day
on
the
basis
of
histopathologic
changes
in
the
liver.

Ten
reproductive
and/
or
developmental
toxicity
studies
(
Ruddick
et
al.,
1983;
Klinefelter
et
al.,
1995;
Narotsky
et
al.,
1997;
NTP,
1998;
Bielmeier
et
al.
2001;
CCC,
2000a,
b,
c,
d;
CCC,
2002)
were
considered
for
derivation
of
the
RfD
in
addition
to
the
chronic
studies.
The
investigations
of
Ruddick
et
al.
(
1983),
Klinefelter
et
al.
(
1995),
and
Narotsky
et
al.
(
1997)
identified
NOAELs
or
LOAELs
in
rats
that
were
substantially
higher
than
the
LOAEL
identified
by
Aida
et
al.
(
1992b)
(
Table
VIII­
4
below).
The
studies
conducted
by
NTP
(
1998)
did
not
observe
developmental
or
reproductive
effects
at
doses
up
to
116
mg/
kg­
day.
The
study
by
Bielmeier
et
al.
(
2001)
identified
a
free­
standing
LOAEL
of
75
mg/
kg­
day
in
F344
rats.
The
studies
conducted
by
CCC
(
2000a,
b)
identified
NOAELs
of
55
and
76
mg/
kg­
day,
respectively,
Draft
­
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Not
Cite
or
Quote
February
20,
2003
VIII
­
21
for
developmental
effects
in
New
Zealand
White
rabbits.
The
study
conducted
in
rats
by
CCC
(
2000c)
identified
concentration­
based
NOAEL
and
LOAEL
values
of
50
ppm
and
150
ppm
for
reduced
body
weight
and
body
weight
gain
in
F
1
pups.
The
study
conducted
in
rats
by
CCC
(
2000d)
identified
NOAEL
and
LOAEL
values
of
45
mg/
kg­
day
and
82
mg/
kg­
day,
respectively
on
the
basis
of
decreased
ossification
sites
per
fetus
per
litter
in
the
forelimb
and
hindlimb.
Lowrange
LOAEL
values
for
maternal
toxicity
ranged
from
13
to
25
mg/
kg­
day
(
CCC,
2000b;
CCC
2000d;
Narotsky
et
al.,
1997).
CCC
(
2002)
identified
a
LOAEL
of
150
ppm
(
approximately
11.6
to
40.2
mg/
kg­
day)
for
delayed
sexual
maturation
in
F
1
male
rats
in
a
two­
generation
study
of
bromodichloromethane
administered
in
drinking
water.
The
LOAEL
for
parental
effects
in
the
study
was
also
150
ppm,
based
on
decreased
body
weight
and
body
weight
gain
in
F
1
females
and
F
1
males
and
females.
Since
these
studies
identified
NOAEL
and/
or
LOAEL
values
substantially
higher
than
that
identified
by
Aida
et
al.
(
1992b),
they
were
not
further
considered
for
derivation
of
the
RfD.

Method
of
Analysis
Selected
data
from
the
candidate
studies
were
analyzed
using
the
benchmark
dose
(
BMD)
modeling
approach.
Initially,
data
sets
for
potentially
sensitive
endpoints
were
selected
as
described
in
U.
S.
EPA
(
1998b)
and
analyzed
using
the
Crump
Benchmark
Dose
Modeling
Software
(
K.
S.
Crump,
Inc.).
Results
of
this
analysis
are
summarized
in
Table
VIII­
5.
Following
the
release
of
Version
1.2
of
the
BMDS
program
(
U.
S.
EPA,
2000a),
a
subset
of
the
most
sensitive
endpoints
identified
using
the
Crump
software
was
reanalyzed
in
accordance
with
proposed
U.
S.
EPA
(
2000b)
recommendations.
An
advantage
of
analysis
with
the
BMDS
software
is
that
several
additional
models
are
available
to
fit
the
data.
The
results
of
the
analysis
with
the
BMDS
software
are
included
in
Table
VIII­
4.

Choice
of
Principal
Study
and
Critical
Effect
for
the
RfD
Three
data
sets
for
histopathological
effects
in
liver
were
analyzed
using
the
BMDS
software
(
Table
VIII­
4).
BMD
modeling
identified
several
endpoints
with
BMD
values
lower
than
the
conventionally
determined
LOAEL
of
6
mg/
kg­
day
(
Aida,
1992b).
The
lowest
BMD
values
were
obtained
for
fatty
degeneration
(
1.9
mg/
kg­
day)
in
male
rats
(
Aida
et
al.,
1992b)
and
for
kidney
cytomegaly
(
2.0
mg/
kg­
day)
in
male
mice
(
NTP,
1987).
Comparably
low
BMD
values
were
also
obtained
for
granulomas
observed
in
the
liver
of
male
rats
(
2.1
mg/
kg­
day)
and
for
fatty
degeneration
in
the
liver
of
female
rats
(
3.1
mg/
kg­
day)
in
the
study
conducted
by
Aida
et
al.
(
1992b).
In
contrast,
the
BMD
values
calculated
for
endpoints
examined
in
other
studies
were
approximately
10­
to
20­
fold
higher.
Draft
­
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Not
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or
Quote
February
20,
2003
VIII
­
22
Table
VIII­
4
Summary
of
Candidate
Studies
for
Derivation
of
the
RfD
for
Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
a
NTP
(
1987)
Rat
F344/
N
M,
F
50
0
50
100
Gavage
(
oil)
102
weeks
(
5
d/
wk)
Body
weight,

clinical
signs,

gross
necropsy,

histology
­­
50
(
lesions
of
kidney
and
liver)
­­
36.5c
(
liver
necrosis
in
male
rats)

NTP
(
1987)
Mouse
B6C3F
1
M,
F
50
0
25
50
Gavage
(
oil)
102
weeks
(
5
d/
wk)
Body
weight,

clinical
signs,

gross
necropsy,

histology
­­
25
(
lesions
of
liver,

kidney,
and
thyroid)
2.0
1.5
(
kidney
cytomegaly
in
male
mice)

Aida
et
al.

(
1992b)
Rat
Wistar
M,
F
40
Male
0
6
26
138
Female
0
8
32
168
Diet
24
months
Body
weight,

clinical
signs,

serum
biochemistry,

gross
necropsy,

histology
­­
6
(
liver
fatty
degeneration
and
granuloma)
3.1
2.1
(
fatty
degeneration
in
liver
of
females
)

1.9
0.8
(
fatty
degeneration
in
liver
of
males)

2.1
1.4
(
Granulomas
in
livers
of
males)

Ruddick
et
al.
(
1983)
b
Rat
SD
F
9­
14
0
ppm
50
ppm
150
ppm
450
ppm
1350
ppm
Gavage
(
oil)
Gestation
days
6­
15
Body
and
organ
weights;
maternal
serum
chemistry;
hematology,
and
histopathology;

developmental
parameters
100
200
(
sternebral
variations)
27
15
(
sternebral
variations)

Narotsky
et
al.
(
1997)
b
Rat
F344
13­

14
0
75
100
Gavage
(
oil)

(
water)
Gestation
days
6­
15
Body
weight,

clinical
signs,
developmental
parameters
25
(
developmental)
50
(
full­
litter
resorption)
48
30
(
full
litter
resorption)
Table
VIII­
4
(
cont.)

Draft
­
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Not
Cite
or
Quote
February
20,
2003
VIII
­
23
Klinefelter
et
al.
(
1995)
b
Rat
F344
M
7
0
22
39
Drinking
water
52
weeks
Body
and
organ
weights,
gross
necropsy,
histology,
sperm
motion
parameters
22
39
(
decreased
sperm
velocities)
­­
­­
d
Bielmeier
et
al.
(
2001)
b
Rat
F344
F
8­
11
0
75
100
Gavage
(
aq)
Gestation
day
9
Full
litter
resorption,

hormone
profiles,
body
weight
­­
75
(
full­
litter
resorption)
23
4.2
(
full­
litter
resorption)

CCC
(
2000a)
b
Rabbit
NZW
F
5
0
4.9
13.9
32.3
76.3
Drinking
Water
Gestation
days
6­
29
Body
wt.,

clinical
signs,

reproductive
developmental
parameters
76
­­
­­
­­
e
CCC
(
2000b)
b
Rabbit
NZW
F
25
0
1.4
13
36
55
Drinking
water
Gestation
days
6­
29
Maternal
feed
and
water
intake,
body
wt.;
gross
lesions;
uterine
weight,
no.
implantation
sites,
uterine
contents,
and
no.

corpora;
Fetal
wt.,
gross
ext.

alterations,
skel.

alterations,
sex,

visceral
alterations
55
(
developmental)
­­
­­
­­
e
13.4
(
maternal)
35.3
(
reduced
maternal
body
weight
gain)
50
(
maternal)
BMDS
software
failed
CCC
(
2000c)
b
Rat
SD
M,
F
10
0
ppm
50
ppm
150
ppm
450
ppm
1350
ppm
Drinking
water
Males
64
days
Females
74
days
Reproductive/
developmental
parameters
50
ppm
150
ppm
­­
­­
Table
VIII­
4
(
cont.)

Draft
­
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Not
Cite
or
Quote
February
20,
2003
VIII
­
24
CCC
(
2000d)
b
Rat
SD
F
25
0.0
2.2
18.4
45.0
82.0
Drinking
water
Gestation
days
6­
21
Reproductive/
developmental
parameters
45
(
developmental)
82
(
reduced
no.
of
ossification
sites
in
phalanges
or
metatarsals
occurring
with
maternal
toxicity)
 
(
developmental)
 
e
18.4
(
maternal)
45
(
reduced
maternal
body
weight
gain)
23
(
maternal)
18
(
reduced
maternal
body
weight
gain)

CCC
(
2002)
Rat
SD
M,
F
30
0
ppm
50
ppm
150
ppm
450
ppm
Drinking
water
Two
generations
Reproductive/
developmental
parameters
50
(
offspring)
150
(
delayed
sexual
maturation
in
F
1
males)
Not
modeled
­­

50
(
parental)
150
(
decrecreased
body
weight
and
body
weight
gain
in
F
0
females
and
F
1
males
and
females)
Not
modeled
­­

a
BMDL10
values
were
derived
using
duration­
adjusted
doses.

b
Ruddick
et
al.
(
1983);
Klinefelter
et
al.
(
1995),
Narotsky
et
al.
(
1997),
Bielmeier
et
al.
(
2001),
and
CCC
(
2000a­
d)
are
included
in
this
table
because
they
are
reproductive
and/
or
developmental
studies.
The
NOAEL,
LOAEL,
BMD,
and
BMDL10
values
listed
are
for
reproductive
and/
or
developmental
endpoints.

c
Data
modeled
using
Crump
BMD
software
d
No
histopathological
abnormalities
were
noted
in
this
study,
and
similar
effects
on
sperm
velocity
were
not
observed
in
NTP
(
1998);
therefore,
data
for
sperm
velocity
were
not
modeled
e
Data
were
not
modeled
since
effects
occurred
at
higher
doses
than
other
candidate
endpoints
 
Indicates
that
data
were
not
modeled
Abbreviations:
NZW,
New
Zealand
White;
SD,
Sprague­
Dawley
NOTE:
The
short­
term
reproductive
and
developmental
toxicity
study
conducted
by
NTP
(
1998)
was
not
included
in
this
table
because
no
developmental
or
reproductive
effects
were
noted
at
dose
levels
ranging
from
67
to
126
mg/
kg­
day.
Draft
­
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or
Quote
February
20,
2003
VIII
­
25
Table
VIII­
5
Summary
of
Preliminary
BMD
Modeling
Results
for
the
Bromodichloromethane
RfD
Study
Endpoint
Modeled
BMDL10
(
mg/
kg­
day)
*

Subchronic
NTP
(
1987)
mouse
study
Focal
necrosis
of
renal
tubular
epithelium
in
males
34
Vacuolated
hepatocytes
in
females
64
Chronic
NTP
(
1987)
rat
study
Kidney
cytomegaly
in
males
No
acceptable
fit
Liver
necrosis
in
males
36.5
Liver
fatty
metamorphosis
in
males
No
acceptable
fit
Clear
cell
changes
in
liver
of
females
No
acceptable
fit
Chronic
NTP
(
1987)
mouse
study
Kidney
cytomegaly
in
males
0.96
Liver
fatty
metamorphosis
in
males
7.5
Thyroid
follicular
cell
hyperplasia
in
females
15
Chronic
Aida
et
al.
(
1992b)
rat
study
Fatty
degeneration
in
liver
of
males
2.38
Granulomas
in
liver
of
males
4.5
Fatty
degeneration
in
liver
of
females
1.20
Granulomas
in
liver
of
females
No
acceptable
fit
*
BMD
modeling
conducted
on
duration­
adjusted
doses
using
the
Crump
BMD
Software
(
K.
S.
Crump,
Inc.).
Draft
­
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Not
Cite
or
Quote
February
20,
2003
VIII
­
26
The
chronic
study
conducted
by
Aida
et
al.
(
1992b)
was
selected
for
derivation
of
the
RfD.
The
lowest
dose
utilized
in
this
study
was
6
mg/
kg­
day
(
in
contrast
to
low
doses
of
22
to
75
mg/
kg­
day
utilized
in
other
candidate
studies),
which
provides
some
information
on
the
shape
of
the
dose­
response
curve
in
the
region
of
interest.
The
lowest
BMD
(
1.9
mg/
kg­
day)
was
obtained
for
fatty
degeneration
in
the
liver
of
male
mice.
The
corresponding
BMDL
10
for
this
endpoint
was
0.8
mg/
kg­
day.
The
incidence
of
this
lesion
was
strongly
dose­
dependent,
with
incidences
of
0/
24,
5/
14,
12/
13,
and
19/
19
observed
at
the
doses
of
0,
6,
25,
and
138
mg/
kg­
day,
respectively.
The
occurrence
of
this
lesion
in
rats
treated
with
bromodichloromethane
is
consistent
with
current
understanding
of
the
mode
of
action
of
brominated
trihalomethanes.
This
endpoint
was
therefore
selected
to
derive
the
RfD
for
bromodichloromethane.

Derivation
of
the
RfD
The
BMDL
10
calculated
for
fatty
degeneration
in
the
liver
of
male
rats
in
the
chronic
rat
study
conducted
by
Aida
et
al.
(
1992b)
was
selected
as
the
most
appropriate
basis
for
derivation
of
the
RfD
for
bromodichloromethane.
The
RfD
is
calculated
according
to
the
following
equation:

RfD
=
(
0.8
mg/
kg­
day)
=
0.003
mg/
kg­
day
(
3
µ
g/
L)
(
300)

where:

0.8
mg/
kg­
day
=
Duration­
adjusted
BMDL
10
based
on
fatty
degeneration
of
the
liver
in
male
rats
300
=
Uncertainty
factor
based
on
NAS/
OW
guidelines.
This
value
includes
a
factor
of
10
to
account
for
intrahuman
variability,
and
a
factor
of
10
for
interspecies
variability,
and
a
factor
of
3
to
account
for
database
deficiencies,
including
lack
of
a
multigeneration
reproductive
toxicity
study
and
database
limitations
and
uncertainty
related
to
possible
human
reproductive
effects
suggested
(
causality
can
not
be
established
from
available
data)
by
epidemiological
studies.

A
composite
UF
of
300
was
used.
The
standard
factors
of
10
were
used
for
interspecies
extrapolation
and
for
protection
of
sensitive
subpopulations.
An
additional
factor
of
3
was
used
to
account
for
database
deficiencies
related
to
possible
reproductive
or
developmental
effects
in
humans.
Use
of
an
additional
uncertainty
factor
of
3
is
supported
by
findings
in
epidemiological
studies
(
Waller
et
al.,
1998;
King
et
al.,
2000)
which
suggest
potential
associations
between
bromodichloromethane
exposure
via
drinking
water
and
adverse
pregnancy
outcomes.
Although
the
results
of
the
epidemiological
studies
can
not
establish
that
bromodichloromethane
caused
the
observed
effects,
they
do
raise
significant
concern
for
potential
reproductive
effects
in
exposed
Draft
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VIII
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27
human
populations,
and
the
inclusion
of
an
additional
uncertainty
factor
is
thus
considered
appropriate
for
protection
of
human
health.

The
DWEL
for
bromodichloromethane
is
calculated
as
follows:

DWEL
=
(
0.003
mg/
kg­
day)
(
70
kg)
=
0.100
mg/
L
(
100
µ
g/
L)
2
L/
day
where:

0.003
mg/
kg­
day
=
RfD
for
bromodichloromethane
70
kg
=
Assumed
weight
of
an
adult
2
L/
day
=
Assumed
water
consumption
by
a
70­
kg
adult
Lifetime
Health
Advisory
The
Lifetime
Health
Advisory
(
HA)
represents
that
portion
of
an
individual's
total
exposure
that
is
attributed
to
drinking
water
and
is
considered
protective
of
noncarcinogenic
health
effects
over
a
lifetime
of
exposure.
Bromodichloromethane
has
been
categorized
with
respect
to
carcinogenic
potential
as
Group
B2:
Probable
human
carcinogen
(
IRIS,
1993a).
Therefore,
in
accordance
with
U.
S.
EPA
Policy,
a
Lifetime
HA
is
not
recommended.

Alternative
Approach
for
Derivation
of
the
RfD
Use
of
the
conventional
NOAEL/
LOAEL
approach
represents
an
alternative
means
for
deriving
the
RfD
and
DWEL.
Aida
et
al.
(
1992b)
identified
a
LOAEL
of
6
mg/
kg­
day
in
male
rats
on
the
basis
of
histopathological
changes
in
the
liver.
Using
this
value
and
a
composite
uncertainty
factor
of
3,000
(
including
factors
of
10
for
interspecies
extrapolation,
protection
of
sensitive
subpopulations,
and
use
of
a
LOAEL,
and
a
factor
of
3
for
database
limitations
and
uncertainty
regarding
potential
reproductive
effects
in
humans),
the
RfD
derived
using
the
conventional
approach
is
0.002
mg/
kg­
day.
Assuming
a
body
weight
of
70
kg
and
drinking
water
ingestion
of
2
L/
day,
the
corresponding
DWEL
is
0.07
mg/
L
(
70
µ
g/
L).
Draft
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2003
VIII
­
28
2.
Carcinogenic
Effects
a.
Categorization
of
Carcinogenic
Potential
Previous
Evaluations
The
Carcinogenic
Risk
Assessment
Verification
Endeavor
(
CRAVE)
group
of
the
U.
S.
EPA
reviewed
the
available
evidence
on
the
carcinogenicity
of
bromodichloromethane
and
assigned
it
to
Group
B2:
probable
human
carcinogen
(
IRIS,
1993a).
Assignment
to
this
category
is
appropriate
for
chemicals
where
there
are
no
or
inadequate
human
data,
but
which
have
sufficient
animal
data
to
indicate
carcinogenic
potential.

IARC
has
recently
re­
evaluated
the
carcinogenic
potential
of
bromodichloromethane
(
IARC
1999a).
IARC
concluded
that
there
is
sufficient
evidence
of
carcinogenicity
for
bromodichloromethane
in
experimental
animals,
but
inadequate
evidence
in
humans.
Thus,
IARC
classified
bromodichloromethane
as
a
Group
2B
carcinogen:
possibly
carcinogenic
to
humans.

Categorization
of
Carcinogenic
Potential
Under
the
Proposed
1999
Cancer
Guidelines
Cancer
Hazard
Summary
Under
the
proposed
guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
bromodichloromethane
is
likely
to
be
carcinogenic
to
humans
by
all
routes.
This
descriptor
is
appropriate
when
the
available
tumor
data
and
other
key
data
are
adequate
to
demonstrate
carcinogenic
potential
to
humans.
This
finding
is
based
on
the
weight
of
experimental
evidence
in
animal
models
which
shows
carcinogenicity
by
modes
of
action
that
are
relevant
to
humans.

Supporting
Information
for
Cancer
Hazard
Assessment
Human
Data
The
information
on
the
carcinogenicity
of
bromodichloromethane
from
human
studies
is
inadequate.
There
are
no
epidemiological
data
specifically
relating
increased
incidence
of
cancer
to
exposure
to
bromodichloromethane.
There
are
equivocal
epidemiological
data
describing
a
weak
association
of
chlorinated
drinking
water
exposures
with
increased
incidences
of
bladder,
rectal,
and
colon
cancer.
U.
S.
EPA
has
determined
that
these
studies
cannot
attribute
the
observed
effects
to
a
single
compound,
as
chlorinated
water
contains
numerous
other
disinfection
byproducts
that
are
potentially
carcinogenic.

Animal
Data
The
carcinogenicity
of
bromodichloromethane
in
male
and
female
animals
has
been
investigated
in
a
well­
designed
and
conducted
corn
oil
gavage
study
conducted
in
rats
and
mice,
a
dietary
exposure
study
in
rats,
and
two
drinking
water
studies
in
rats.
Additional
data
are
available
from
a
study
in
which
male
Strain
A
mice
were
exposed
to
bromodichloromethane
by
Draft
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February
20,
2003
VIII
­
29
intraperitoneal
injection.
No
data
are
available
on
the
carcinogenic
potential
of
bromodichloromethane
administered
via
the
inhalation
or
dermal
routes.

In
the
corn
oil
gavage
study
(
NTP,
1987),
statistically
significant
increases
were
observed
in
the
incidences
of
neoplasms
of
the
large
intestine
and
kidney
in
male
and
female
rats,
the
kidney
in
male
mice,
and
liver
in
female
mice.
The
neoplasms
observed
in
the
large
intestine
and
kidney
are
considered
rare
neoplasms
based
on
historical
control
data
for
the
tested
strains.
In
the
feeding
study
(
Aida
et
al.,
1992b),
exposure
to
microencapsulated
bromodichloromethane
did
not
result
in
statistically
significant
increases
in
any
tumor
type.
Observed
neoplastic
lesions
included
three
cholangiocarcinomas
and
two
hepatocellular
adenomas
in
high
dose
females,
one
hepatocellular
adenoma
in
a
control
female,
one
cholangiosarcoma
in
a
high
dose
male,
and
one
hepatocellular
carcinoma
each
in
a
low­
and
a
high
dose
male.
In
the
drinking
water
study
conducted
by
Tumasonis
et
al.
(
1985),
hepatic
neoplastic
nodules,
hepatic
adenofibrosis,
and
lymphosarcoma
were
significantly
increased
in
female
rats.
No
significant
increase
in
the
occurrence
of
any
tumor
type
was
observed
in
male
rats.
Renal
adenoma
or
adenocarcinoma
were
noted
in
two
males
and
one
female
treated
with
bromodichloromethane,
while
neither
tumor
type
was
reported
in
the
control
group.
In
the
drinking
water
study
conducted
by
George
et
al.
(
2002),
the
prevalence
of
neoplastic
lesions
in
the
liver
was
significantly
increased
only
at
the
lowest
administered
dose.
Intraperitoneal
injection
of
Strain
A
mice
with
bromodichloromethane
resulted
in
an
apparent
increase
in
the
number
of
pulmonary
adenomas
per
animals,
although
the
response
did
not
reach
statistical
significance
in
any
dose
group.

Structural
Analogue
Data
Trihalomethanes
structurally
related
to
bromodichloromethane
have
shown
varying
degrees
of
carcinogenic
potential
in
rodents.
Chloroform,
the
most
extensively
characterized
trihalomethane,
is
reported
to
be
carcinogenic
at
high
doses
in
several
chronic
animal
bioassays,
with
significant
increases
in
the
incidence
of
liver
tumors
in
male
and
female
mice
and
significant
increases
in
the
incidence
of
kidney
tumors
in
male
rats
and
mice
(
U.
S.
EPA,
2001).
The
occurrence
of
tumors
in
animals
exposed
to
chloroform
is
demonstrably
species­,
strain­,
and
gender­
specific,
and
has
only
been
observed
under
dose
conditions
that
caused
cytotoxicity
and
regenerative
cell
proliferation
in
the
target
organ.
The
cancer
database
for
structurally­
related
brominated
trihalomethanes
is
more
limited,
but
includes
well­
conducted
studies
performed
by
the
National
Toxicology
Program.
In
a
two­
year
corn
oil
gavage
study
of
bromoform,
NTP
(
1989a)
found
clear
evidence
for
carcinogenicity
in
female
rats
and
some
evidence
of
carcinogenicity
in
male
rats
based
on
occurrence
of
tumors
of
the
large
intestine
(
adenomatous
polyps
or
adenocarcinoma).
In
a
two­
year
corn
oil
gavage
study
of
dibromochloromethane,
NTP
(
1985)
determined
that
there
was
some
evidence
of
carcinogenicity
in
female
mice
and
equivocal
evidence
of
carcinogenicity
in
male
mice,
based
on
the
occurrence
of
hepatocellular
adenomas
and
carcinomas.
Other,
less
well­
documented,
oral
exposure
studies
(
Tobe
et
al.,
1982;
Kurokawa,
1987;
Voronin
et
al.,
1987)
found
no
evidence
for
carcinogenicity
of
bromoform
or
dibromochloromethane.

Other
Key
Data
Draft
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VIII
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30
Bromodichloromethane
is
formed
as
a
byproduct
of
drinking
water
disinfection
with
chlorine.
Exposure
to
bromodichloromethane
may
occur
via
ingestion
of
tap
water,
via
dermal
contact
during
showering
or
bathing,
or
by
inhalation
of
bromodichloromethane
volatilized
during
household
activities.
Absorption
of
single
oral
doses
appears
to
be
extensive.
Bromodichloromethane
is
rapidly
metabolized
and
eliminated
predominately
as
expired
volatiles,
CO
2,
or
CO.
Only
a
small
amount
(
less
than
10%)
is
eliminated
in
urine
or
in
feces.
No
comprehensive
tissue
data
are
available
regarding
the
bioaccumulation
or
retention
of
bromodichloromethane
following
repeated
exposure.
However,
because
of
the
rapid
metabolism
and
excretion
of
bromodichloromethane,
marked
accumulation
and
retention
is
not
expected.

Bromodichloromethane
itself
is
not
directly
reactive
with
DNA.
Metabolism
to
reactive
species
is
a
prerequisite
for
toxicity,
as
inferred
from
metabolic
induction
and
inhibition
studies.
In
vitro
and
in
vivo
studies
of
the
mutagenic
and
genotoxic
potential
of
bromodichloromethane
have
yielded
both
positive
and
negative
results.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid)
and,
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
However,
because
a
majority
of
studies
yielded
positive
results,
bromodichloromethane
is
considered
to
be
at
least
weakly
mutagenic
and
genotoxic.
Recent
studies
conducted
with
strains
of
Salmonella
engineered
to
express
rat
theta­
class
glutathione
S­
transferase
suggest
that
mutagenicity
of
the
brominated
trihalomethanes
may
be
mediated
by
glutathione
conjugation.

Mode
of
Action
The
mode
of
action
for
tumor
induction
by
bromodichloromethane
has
not
been
clearly
elucidated
and
may
involve
contributions
from
multiple
bioactivation
pathways.
In
each
case,
toxicity
is
believed
to
result
from
interaction
of
reactive
metabolites
with
cellular
macromolecules.
Proposed
bioactivation
pathways
for
bromodichloromethane
include:
1)
production
of
reactive
dihalocarbonyls
by
oxidative
metabolism;
2)
production
of
reactive
dihalomethyl
radicals
by
oxidative
metabolism;
and
3)
formation
of
DNA­
reactive
species
via
a
glutathione­
dependent
pathway.
The
relative
contribution
of
each
pathway
to
tumor
induction
by
bromodichloromethane
has
not
been
characterized.
It
is
possible
that
only
the
latter
two
processes
lead
to
DNA
damage
in
vivo,
because
the
highly
reactive
dihalocarbonyl
intermediate
may
not
survive
long
enough
to
enter
the
nucleus
and
react
with
DNA.
For
this
reason,
cytotoxicity
may
be
the
primary
consequence
of
the
oxidative
pathway.
Cytotoxicity
coupled
with
regenerative
hyperplasia
is
considered
the
primary
mode
of
action
for
tumor
formation
following
exposure
to
high
concentrations
of
chloroform,
a
structurally­
related
trihalomethane
which
has
low
genotoxic
potential.
Data
to
support
a
similar
primary
mode
of
action
for
tumor
development
in
liver,
kidney,
and
large
intestine
are
currently
lacking
for
bromodichloromethane.
In
the
absence
of
such
information,
combined
with
a
positive
weight­
of­
evidence
evaluation
for
genotoxicity,
the
mode
of
action
for
tumor
development
is
assumed
to
be
a
linear
process.
The
processes
leading
to
tumor
formation
in
animals
are
expected
to
be
relevant
to
humans.

Conclusion
Draft
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February
20,
2003
VIII
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31
Under
the
proposed
guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
bromodichloromethane
is
likely
to
be
carcinogenic
to
humans
by
the
oral
route.
This
weight­
ofevidence
evaluation
is
based
on
1)
observations
of
tumors
in
animals
treated
by
oral
pathways;
2)
lack
of
epidemiological
data
specific
to
bromodichloromethane
and
equivocal
data
for
drinking
water
drinking
water
exposures
that
cannot
reliably
be
attributed
to
bromodichloromethane
among
multiple
other
disinfection
byproducts;
3)
positive
results
for
a
majority
of
the
available
genotoxicity
and
mutagenicity
tests;
and
4)
metabolism
and
mode
of
action
that
are
reasonably
expected
to
be
comparable
across
species.
Although
no
cancer
data
exist
for
exposures
via
the
dermal
or
inhalation
pathways,
the
weight­
of­
evidence
conclusion
is
considered
to
be
applicable
to
these
pathways
as
well.
The
finding
for
inhalation
is
based
on
the
observation
that
patterns
of
metabolizing
enzyme
activity
in
male
rats
are
similar
for
exposure
via
the
inhalation
and
gavage
routes.
Bromodichloromethane
absorbed
through
the
skin
is
expected
to
be
metabolized
and
cause
toxicity
in
much
the
same
way
as
bromodichloromethane
absorbed
by
the
oral
and
inhalation
routes.

b.
Choice
of
Study
for
Quantification
of
Carcinogenic
Risk
In
accordance
with
the
Proposed
1999
Cancer
Guidelines
(
U.
S.
EPA,
1999),
quantification
of
cancer
risk
is
appropriate
for
compounds
categorized
as
likely
to
be
carcinogenic
to
humans.
Five
oral
exposure
studies
were
available
for
quantification
of
the
carcinogenic
risk
associated
with
exposure
to
bromodichloromethane.
Detailed
summaries
of
these
studies
are
provided
in
Section
V.
G.
1.
The
two­
year
study
conducted
by
NTP
(
1987)
in
rats
and
mice
was
selected
for
quantification
of
carcinogenic
effects
associated
with
exposure
to
bromodichloromethane.
Selection
of
this
study
was
based
on
significantly
increased
incidence
of
several
tumor
types,
monotonic
dose
response
curves,
and
comprehensive
documentation
of
study
design
and
results.

In
the
NTP
(
1987)
study,
groups
of
male
and
female
F344/
N
rats
(
50/
sex/
dose)
received
0,
50,
or
100
mg/
kg­
day
gavage
doses
of
bromodichloromethane
in
corn
oil.
The
doses
were
administered
5
days/
week
for
102
weeks.
In
a
similar
experiment,
groups
of
male
and
female
B6C3F
1
mice
(
50/
sex/
dose)
were
administered
doses
of
0,
25,
or
50
mg/
kg­
day
(
males)
or
0,
75,
or
150
mg/
kg­
day
(
females)
for
5
days/
week
for
102
weeks.
All
animals
were
subjected
to
gross
and
microscopic
examinations
for
neoplastic
lesions.
Survival
of
all
dosed
animals
was
comparable
to
or
greater
than
the
corresponding
control
group.
Statistically
significant
increases
were
observed
in
the
incidences
of
neoplasms
of
the
large
intestine
and
kidney
in
male
and
female
rats,
the
kidney
in
male
mice,
and
the
liver
in
female
mice
(
Table
VIII­
6).
The
authors
noted
that
neoplasms
of
the
large
intestine
and
kidney
are
uncommon
tumors
in
F344/
N
rats
and
B6C3F
1
mice.
They
concluded
that
under
the
conditions
of
these
2­
year
gavage
studies,
clear
evidence
of
carcinogenic
activity
existed
for
male
and
female
rats
and
mice.
Draft
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February
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VIII
­
32
Table
VIII­
6
Tumor
Frequencies
in
Rats
and
Mice
Exposed
to
Bromodichloromethane
in
Corn
Oil
for
2
Years
­
Adapted
from
NTP
(
1987)

Animal
Tissue/
Tumor
Tumor
Frequency
Male
rat
Control
50
mg/
kg
100
mg/
kg
Large
intestinea
Adenomatous
polyp
0/
50
3/
49
33/
50b
Adenocarcinoma
0/
50
11/
49b
38/
50b
Combined
0/
50
13/
49b
45/
50b
Kidney
a
Tubular
cell
adenoma
0/
50
1/
49
3/
50
Tubular
cell
adenocarcinoma
0/
50
0/
49
10/
50b
Combined
0/
50
1/
49
13/
50b
Female
rat
Control
50
mg/
kg
100
mg/
kg
Large
intestine
c
Adenomatous
polyp
0/
46
0/
50
7/
47b
Adenocarcinoma
0/
46
0/
50
6/
47b
Combined
0/
46
0/
50
12/
47b
Kidney
Tubular
cell
adenoma
0/
50
1/
50
6/
50b
Tubular
cell
adenocarcinoma
0/
50
0/
50
9/
50b
Combined
0/
50
1/
50
15/
50b
Male
mouse
Control
25
mg/
kg
50
mg/
kg
Kidney
d
Tubular
cell
adenoma
1/
46
2/
49
6/
50
Tubular
cell
adenocarcinoma
0/
46
0/
49
4/
50
Combined
1/
46
2/
49
9/
50b
Female
mouse
Control
75
mg/
kg
150
mg/
kg
Liver
Hepatocellular
adenoma
1/
50
13/
48b
23/
50b
Hepatocellular
carcinoma
2/
50
5/
48
10/
50b
Combined
3/
50
18/
48b
29/
50b
a
One
rat
died
at
week
33
in
the
low­
dose
group
and
was
eliminated
from
the
cancer
risk
calculation.
b
Statistically
significant
at
p<
0.05,
compared
to
controls.
c
Intestine
not
examined
in
four
rats
from
control
group
and
three
rats
from
high­
dose
group.
d
In
the
control
group,
two
mice
died
during
the
first
week,
one
mouse
died
during
week,
nine
and
one
escaped
in
week
79.
One
mouse
in
the
low­
dose
group
died
in
the
first
week.
All
of
these
mice
were
eliminated
from
the
cancer
risk
calculations.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
33
Use
of
the
NTP
rodent
studies
(
NTP,
1987)
for
derivation
of
cancer
risk
estimates
for
bromodichloromethane
is
complicated
by
the
use
of
a
corn
oil
as
a
dosing
vehicle.
Although
a
vehicle
effect
has
not
been
reported
for
brominated
trihalomethanes,
it
can
be
inferred
from
studies
of
chloroform
carcinogenicity
that
such
an
effect
might
exist,
at
least
for
hepatic
tumors
in
mice.
Therefore,
in
the
case
of
bromodichloromethane,
the
U.
S.
EPA
believes
that
the
most
appropriate
basis
of
the
cancer
risk
estimate
is
the
incidence
of
renal
tumors
in
male
mice.
Renal
tumors
are
considered
to
be
appropriate
because
these
tumors
were
increased
in
a
dose­
dependent
manner
in
both
mice
(
male)
and
rats
(
both
sexes).

c.
Choice
of
Approach
and
Rationale
The
LMS
model
(
U.
S.
EPA,
1986)
and
the
default
linear
approach
described
by
the
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1996;
1999)
were
used
to
quantify
the
cancer
risk
associated
with
exposure
to
bromodichloromethane.
Although
data
are
mixed,
the
weight
of
evidence
indicates
that
bromodichloromethane
is
mutagenic
(
see
Section
V.
F.
1).
Furthermore,
recent
studies
(
Melnick,
et
al.
1998;
George
et
al.,
2002)
suggest
that
induction
of
hepatic
tumors
occurs
at
doses
of
bromodichloromethane
that
have
marginal
or
no
effect
on
hepatocyte
labeling
index,
indicating
that
regenerative
hyperplasia
is
not
required
for
tumor
induction.
Thus,
use
of
a
linear
approach
was
considered
appropriate
for
quantification
of
cancer
risk
associated
with
exposure
to
bromodichloromethane.

d.
Cancer
Potency
and
Risk
Estimates
The
available
estimates
for
cancer
risk
associated
with
bromodichloromethane
are
summarized
in
Table
VIII­
7.
U.
S.
EPA
(
1994b)
recommended
use
of
a
cancer
potency
estimate
of
6.2
x
10­
2
(
mg/
kg­
day)­
1
as
reported
in
IRIS
(
1993a).
This
value
was
derived
in
accordance
with
the
1986
Guidelines
for
Carcinogenic
Risk
Assessment
(
U.
S.
EPA,
1986),
based
on
the
occurrence
of
renal
tumors
in
male
mice.
A
unit
risk
of
1.8
x
10­
6
was
estimated
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
approximately
6
µ
g/
L
associated
with
a
10­
5
risk
(
0.6
µ
g/
L
for
10­
6
risk).

A
cancer
potency
value
of
3.5
x
10­
2
(
mg/
kg­
day)­
1
(
U.
S.
EPA,
1998b)
was
derived
using
the
LMS
method
and
an
animal­
to­
human
conversion
factor
of
body
weight3/
4
(
Table
VIII­
7).
The
use
of
body
weight3/
4
is
consistent
with
recommendations
in
U.
S.
EPA
(
1992b).
This
potency
factor
is
also
based
on
the
occurrence
of
renal
tumors
in
male
mice.
A
unit
risk
of
1
x
10­
6
(
µ
g/
L)­
1
was
estimated
for
bromodichloromethane
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
approximately
10
µ
g/
L
associated
with
a
10­
5
risk
(
1
µ
g/
L
for
10­
6
risk).

Cancer
risk
estimates
were
also
obtained
using
the
LED
10
(
the
lower
95%
confidence
limit
on
a
dose
associated
with
10%
extra
risk)
for
renal
tumors
in
mice
and
assuming
a
linear
mode
of
action
for
the
carcinogenicity
of
bromodichloromethane
(
Table
VIII­
7).
A
cancer
potency
value
of
3.4
x
10­
2
(
mg/
kg­
day)­
1
was
derived
using
this
approach.
A
unit
risk
of
9.6
x
10­
7
(
µ
g/
L)­
1
was
estimated
for
bromodichloromethane
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
34
approximately
10
µ
g/
L
associated
with
a
10­
5
risk
(
1
µ
g/
L
for
10­
6
risk).
These
values
are
closely
similar
to
corresponding
values
derived
using
the
LMS
approach
with
body
weight
scaling
to
the
3/
4
power.

Table
VIII­
7
Summary
of
Cancer
Risk
Estimates
for
Bromodichloromethane
Method
of
Estimation
Tumor
Site
Species
Sex
Slope
Factor
(
mg/
kg­
day)­
1
Unit
Risk
(
µ
g/
L)­
1
LED
10
(
µ
g/
kgday
10­
5
Risk
Concentration
(
µ
g/
L)

LMS
Method
Using
BW3/
4Conversion
(
U.
S.
EPA,
1998b)
Liver
Mouse
F
6.9
×
10­
2
2.0
×
10­
6
­
5
Kidney
Rat
M
F
5.5
×
10­
3
6.1
×
10­
3
1.6
×
10­
7
1.7
×
10­
7
­
64
57
Mouse
M
3.5
×
10­
2
1.0
×
10­
6
­
10
Large
intestine
Rat
M
F
1.7
×
10­
2
6.1
×
10­
3
4.9
×
10­
7
1.7
×
10­
7
­
20
57
LMS
Method
Using
BW2/
3Conversion
U.
S.
EPA
(
1994b)*
Liver
Mouse
F
1.3
×
10­
1
3.7
×
10­
6
­
3
Kidney
Rat
M
F
8.7
×
10­
3
9.5
×
10­
3
2.5
×
10­
7
2.7
×
10­
7
­
40
37
Mouse
M
6.2
×
10­
2
1.8
×
10­
6
­
6
Large
intestine
Rat
M
F
2.5
×
10­
2
4.9
×
10­
3
7.1
×
10­
7
1.4
×
10­
7
­
14
72
LED
10/
Linear
Method
(
U.
S.
EPA,
1998b)
Liver
Mouse
F
6.5
×
10­
2
1.9x10­
6
1.5
×
103
5
Kidney
Rat
M
F
8.1
×
10­
3
8.8
×
10­
3
2.3x10­
7
2.5x10­
7
1.2
×
104
1.1
×
104
43
40
Mouse
M
3.4
x
10­
2
9.6
x10­
7
3.0
x103
10
Large
intestine
Rat
M
F
2.8
×
10­
2
1.0
×
10­
2
8x10­
7
3x10­
7
3.5
×
103
9.6
×
103
12
34
*
Based
on
information
adapted
from
IRIS
(
1993a)
Draft
­
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or
Quote
February
20,
2003
VIII
­
35
B.
Dibromochloromethane
1.
Noncarcinogenic
effects
a.
One­
day
Health
Advisory
Four
candidate
studies
that
investigated
the
acute
oral
toxicity
of
dibromochloromethane
were
available.
These
studies
are
summarized
in
Table
VIII­
8
(
below).
Bowman
et
al.
(
1978)
administered
dibromochloromethane
by
gavage
to
ICR
Swiss
mice
at
doses
ranging
from
500
to
4,000
mg/
kg
and
found
that
sedation
and
anesthesia
resulted
at
doses
of
500
mg/
kg
or
higher.
NTP
(
1985)
conducted
a
preliminary
range­
finding
study
in
which
F344/
N
rats
and
B6C3F
1
mice
were
administered
dibromochloromethane
by
gavage
at
doses
ranging
from
160
to
2,500
mg/
kg
and
found
that
death
may
result
from
doses
at
310
mg/
kg
or
higher
in
mice
or
rats.
More
recently,
Müller
et
al.
(
1997)
investigated
the
cardiotoxic
effects
of
acute
oral
dibromochloromethane
exposure
in
male
Wistar
rats.
In
this
study,
rats
administered
doses
of
83
or
167
mg/
kg
exhibited
transient
changes
in
cardiovascular
parameters,
while
rats
administered
doses
of
333
or
667
mg/
kg
exhibited
persistent
alterations
in
at
least
one
of
the
cardiovascular
parameters
that
lasted
throughout
the
postexposure
observation
period.
Finally,
Korz
and
Gatterman
(
1997)
investigated
the
behavioral
toxicity
of
acute
oral
dibromochloromethane
exposure
in
male
golden
hamsters
and
observed
only
transient
effects
on
the
behavioral
parameters
investigated.

These
studies
were
not
considered
adequate
for
deriving
the
One­
day
HA,
since
evaluated
more
sensitive
endpoints
such
as
histopathology.
Therefore,
the
Ten­
day
HA
for
dibromochloromethane
(
0.6
mg/
L)
calculated
below
is
recommended
for
use
as
the
One­
day
HA.

b.
Ten­
day
Health
Advisory
Studies
considered
for
derivation
of
the
Ten­
day
HA
for
dibromochloromethane
are
summarized
in
Table
VIII­
9
below.
The
key
studies
in
this
group
are
those
of
Aida
et
al.
(
1992a),
Condie
et
al.
(
1983),
and
Melnick
et
al.
(
1998).
These
studies
reported
effects
on
sensitive
endpoints
and
had
data
suitable
for
BMD
analysis.
Use
of
the
remaining
studies
was
limited
by
a
variety
of
considerations,
including
lack
of
data
suitable
for
BMD
analysis
(
Chu
et
al.,
1982a;
NTP,
1996;
Coffin
et
al.,
2000),
toxicological
relevance
or
difficulty
in
interpretation
of
the
most
sensitive
endpoint
(
NTP,
1985;
Munson
et
al.
1982),
and
occurrence
of
effects
only
at
the
frank
toxicity
level
(
NTP,
1985).

Melnick
et
al.
(
1998)
administered
dibromochloromethane
to
female
B6C3F
1
mice
by
gavage
for
5
days/
week
for
3
weeks
and
identified
a
NOAEL
of
100
mg/
kg­
day
(
durationadjusted
NOAEL
of
71
mg/
kg­
day)
and
a
LOAEL
of
192
mg/
kg­
day
(
duration­
adjusted
LOAEL
of
137
mg/
kg­
day)
based
on
histologic
changes
in
the
liver
(
hepatocyte
hydropic
degeneration).
BMD
analysis
calculated
BMD
and
BMDL
10
values
of
112
and
68
mg/
kg­
day,
respectively.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
36
Table
VIII­
8
Summary
of
Candidate
Studies
for
Derivation
of
the
One­
day
HA
for
Dibromochloromethane
Reference
Species
Route
Exposure
Duration
Dose
(
mg/
kg­
day)
Result
Bowman
et
al.

(
1978)
Mouse
ICR
Swiss
Gavage
(
aqueous)
Single
dose
500
­
4000
Sedation;
anesthesia;
increased
mortality
NTP
(
1985)
Rat
F344/
N
Mouse
B6C3F1
Gavage
(
corn
oil)
Single
dose
160
­
2500
Lethargy;
death
Müller
et
al.

(
1997)
Rat
Wistar
Gavage
(
olive
oil)
Single
dose
83
­
667
Transient
changes
in
blood
pressure;
effects
on
cardiac
muscle
contractility
Korz
and
Gatterman
(
1997)
Hamster
Gavage
(
olive
oil)
Single
dose
50
Transient
changes
in
posttreatment
behavior
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
37
Table
VIII­
9
Summary
of
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Dibromochloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Aida
et
al.

(
1992a)
Rat
Wistar
M,
F
7
Males
0
18
56
173
Females
0
34
101
332
Feed
1
month
Body
weight,
clinical
signs,
serum
biochemistry,

hematology,
histology
18.3
56
(
liver
histopathology)
29
14
6.7
(
Liver
cell
vacuolation
in
females)

5.5
(
Liver
cell
vacuolation
in
males)

Chu
et
al.

(
1982a)
Rat
SD
M
10
0
0.7
8.5
68
Drinking
water
28
days
Clinical
signs,
serum
biochemistry,

histology
68
­­
Not
modeled
­­

Condie
et
al.

(
1983)
Mouse
CD­
1
M
8­
16
0
37
74
147
Gavage
(
oil)
14
days
Serum
enzymes,
PAH
uptake
in
vitro,

histology
74
147
(
elevated
ALT,

decreased
PAH,

liver
and
kidney
histopathology)
3.5
11
1.6
(
Renal
mesangial
hypertrophy)

6.9
(
hepatic
cytoplasmic
vacuolation)

Melnick
et
al.

(
1998)
Mouse
B6C3F
1
F
10
0
50/
37
100/
71
192/
137
417/
298
Gavage
(
oil)
3
weeks
(
5
d/
wk)
Body
and
liver
weights,
serum
chemistry,
liver
histology
100
(
marginal)
192
(
liver
histopathology)
112*
68
(
hepatic
hydropic
degeneration)

Munson
et
al.

(
1982)
Mouse
CD­
1
M,
F
8­
12
0
50
125
250
Gavage
(
aqueous)
14
days
Body
and
organ
weights,
serum
chemistry,
hematology,
immune
function
50
125
(
decreased
humoral
immunity)
Not
modeled
­­
Table
VIII­
9
(
cont.)

Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Draft
­
Do
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Cite
or
Quote
February
20,
2003
VIII
­
38
NTP
(
1985)
Rat
F344/
N
M,
F
5
0
60
125
250
500
1000
Gavage
(
oil)
14
days
Body
weight,
clinical
signs,
gross
necropsy
250
500
(
FEL)

(
mortality,

lethargy,
gross
pathology)
Not
modeled
­­

NTP
(
1985)
Mouse
B6C3F
1
M,
F
5
0
30
60
125
250
500
Gavage
(
oil)
14
days
Body
weight,
clinical
signs,
gross
necropsy
60
125
(
stomach
lesions)
143
218
54
(
stomach
nodules
­
males)

77
(
stomach
nodules
­
females)

NTP
(
1996)
Rat
F344/
N
M,
F
10
Males
0
4
12
28
Females
0
6­
7
17­
20
48­
48
Drinking
water
29
days
Body
weight,
serum
chemistry,
hematology,
gross
necropsy,
histology,

sperm
evaluation
28
­­
Not
modeled
 
Coffin
et
al.

(
2000)
Mouse
B6C3F
1
F
10
0
100
300
Gavage
(
oil)
11
days
Relative
liver
wt.;

liver
histopathology;

labeling
index
­­
100
Not
modeled
­­

*
BMD
and
BMDL10
calculated
using
duration­
adjusted
doses
Abbreviations
:
FEL,
Frank
Effect
Level;
SD,
Sprague­
Dawley
 
No
data
Draft
­
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Not
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or
Quote
February
20,
2003
VIII
­
39
These
values
are
considerably
higher
(
approximately
5­
to
10­
fold)
than
BMD
and
BMDL
10
values
calculated
for
hepatic
effects
using
data
from
Condie
et
al.
(
1983)
or
Aida
et
al.
(
1992a).

Condie
et
al.
(
1983)
administered
dibromochloromethane
by
gavage
to
male
CD­
1
mice
for
14
days
and
identified
a
NOAEL
of
74
mg/
kg­
day
and
a
LOAEL
of
148
mg/
kg­
day
based
on
minimal
to
moderate
liver
and
kidney
injury.
Histologic
changes
in
the
liver
included
focal
inflammation
and
cytoplasmic
vacuolization
similar
to
that
observed
in
the
study
by
Aida
et
al.
(
1992a).
Effects
in
the
kidney
included
minimal
to
slight
epithelial
hyperplasia
at
the
high
dose
and
minimal
to
slight
mesangial
hypertrophy
at
all
(
non­
control)
doses.
Data
for
cytoplasmic
vacuolization
and
renal
mesangial
hypertrophy
were
analyzed
by
BMD
modeling.
The
lowest
BMD
and
BMDL
10
(
3.5
and
1.6
mg/
kg­
day,
respectively)
among
all
candidate
studies
were
identified
for
mesangial
hypertrophy.
However,
the
pattern
of
dose­
response
for
this
endpoint
(
0/
16,
4/
5,
7/
10,
7/
10
at
doses
of
0,
37,
74,
and
147
mg/
kg­
day,
respectively)
resulted
in
generally
poor
curve
fits
(
0.1<
goodness
of
fit
p
value>
0.49)
and
a
high
degree
of
model­
dependence
(
See
summary
of
modeling
results
in
Appendix
A).
Thus,
confidence
in
the
reliability
of
the
BMD
for
renal
effects
was
low.
The
BMD
and
BMDL
10
values
for
hepatic
cytoplasmic
vacuolization
were
higher
(
11
and
6.9
mg/
kg­
day,
respectively).
These
results
were
based
on
incidences
of
1/
16,
3/
5,
4/
10,
and
8/
10
at
doses
of
0,
37,
74,
and
147
mg/
kg­
day,
respectively.

The
study
by
Aida
et
al.
(
1992a)
was
selected
as
the
basis
for
derivation
of
the
Ten­
day
HA.
In
this
study,
Wistar
rats
of
both
sexes
were
administered
microencapsulated
dibromochloromethane
in
the
diet
for
one
month.
The
dose
levels
ranged
from
18.3
to
173.3
mg/
kg­
day
for
males
and
from
34.0
to
332.5
mg/
kg­
day
for
females.
A
NOAEL
of
18.3
mg/
kgday
and
a
LOAEL
of
56.2
mg/
kg­
day
were
identified
based
on
histologic
changes
(
cell
vacuolization,
swelling,
and
single
cell
necrosis)
in
the
livers
of
male
rats.
BMD
analysis
of
data
for
hepatic
cell
vacuolization
calculated
BMD
and
BMDL
10
values
of
29
and
6.7
mg/
kg­
day
in
females
and
14
and
5.5
mg/
kg­
day
in
males,
respectively.
The
BDML
10
for
hepatic
cell
vacuolization
in
male
rats
was
selected
for
calculation
of
the
10­
day
HA
because
it
was
considered
the
lowest
reliable
value
based
on
examination
of
the
raw
data
and
modeling
results.
The
incidence
of
this
endpoint
was
0/
7,
1/
7,
3/
7,
and
7/
7
at
doses
of
0,
18,
56,
and
173
mg/
kgday
respectively.

Based
on
the
BMDL
10
calculated
from
the
data
of
Aida
et
al.
(
1992a),
the
Ten­
day
HA
is
derived
as
follows:

Ten­
day
HA
=
(
5.5
mg/
kg­
day)
(
10
kg)
=
0.55
mg/
L
(
rounded
to
0.6
mg/
L)
(
100)
(
1
L/
day)

5.5
mg/
kg­
day
=
BMDL
10
based
on
hepatic
cell
vacuolization
in
rats
fed
dibromochloromethane
for
one
month
10
kg
=
Assumed
body
weight
of
a
child
Draft
­
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or
Quote
February
20,
2003
VIII
­
40
100
=
Composite
uncertainty
factor
based
on
NAS/
OW
guidelines;
includes
a
factor
of
10
for
interspecies
variation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
The
Ten­
day
HA
was
calculated
using
the
conventional
NOAEL/
LOAEL
approach
for
comparison
with
the
value
obtained
using
the
BMD
approach.
The
Aida
et
al.
(
1992a)
study
identified
a
NOAEL
of
18.3
mg/
kg­
day
based
on
the
absence
of
hepatic
effects
in
rats.
Using
this
value
and
and
the
assumptions
described
above,
the
Ten­
day
HA
would
be
1.8
mg/
L
(
rounded
to
2
mg/
L).

c.
Longer­
term
Health
Advisory
Four
candidate
studies
were
considered
for
derivation
of
the
Longer­
term
HA
for
dibromochloromethane.
These
studies
are
summarized
in
Table
VIII­
10
(
below).
Selected
data
from
three
of
these
studies
were
modeled
using
the
BMD
approach.
The
results
of
BMD
analysis
are
included
in
Table
VIII­
11.

Chu
et
al.
(
1982b)
administered
dibromochloromethane
to
Sprague­
Dawley
rats
in
the
drinking
water
at
doses
ranging
from
0.57
to
236
mg/
kg­
day.
This
study
identified
a
NOAEL
of
49
mg/
kg­
day,
and
a
LOAEL
of
224
mg/
kg­
day
based
on
mild
hepatic
lesions
(
increased
cytoplasmic
volume
and
vacuolation
due
to
fatty
infiltration)
observed
in
males.
BMD
and
BMDL
10
values
of
18
and
5.3
mg/
kg­
day,
respectively,
were
calculated
for
hepatic
vacuolization
using
the
BMDS
software.
Daniel
et
al.
(
1990)
identified
a
LOAEL
of
50
mg/
kg­
day
based
on
hepatic
lesions
(
centrilobular
lipidosis)
observed
in
male
Sprague­
Dawley
rats
and
on
kidney
lesions
(
tubular
degeneration)
observed
in
female
Sprague­
Dawley
rats
administered
dibromochloromethane
by
gavage
for
90
consecutive
days.
BMD
and
BMDL
10
values
of
20
and
4.2
mg/
kg­
day,
respectively,
were
calculated
for
renal
tubular
degeneration
using
the
Crump
BMD
software
(
K.
S.
Crump,
Inc.).
NTP
(
1985)
administered
doses
of
dibromochloromethane
ranging
from
15
to
250
mg/
kg­
day
to
male
and
female
mice.
The
compound
was
administered
by
gavage
in
corn
oil,
five
days
per
week
for
13
weeks.
NOAEL
and
LOAEL
values
of
125
and
250
mg/
kg­
day
were
identified
on
the
basis
of
renal
and
hepatic
lesions.
BMD
and
BMDL
10
values
were
not
calculated
since
lesions
occurred
only
at
the
high
dose
of
250
mg/
kg­
day.

The
NTP
(
1985)
study
conducted
in
rats
was
selected
as
the
basis
for
derivation
of
the
Longer­
term
HA.
In
this
study,
F344/
N
rats
were
administered
dibromochloromethane
by
gavage
at
dose
levels
ranging
from
15
to
250
mg/
kg
for
5
days/
week
for
13
weeks.
Severe
lesions
and
necrosis
of
the
kidney,
liver,
and
salivary
glands
were
observed
primarily
at
the
high
dose.
However,
males
exhibited
a
dose­
dependent
increase
in
the
frequency
of
clear
cytoplasmic
vacuoles
indicative
of
fatty
metamorphosis
in
the
liver.
This
effect
reached
statistical
Draft
­
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February
20,
2003
VIII
­
41
Table
VIII­
10
Summary
of
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
for
Dibromochloromethane
Reference
Species
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)*
BMDL
10
(
mg/
kg­
day)
*

Chu
et
al.

(
1982b)
Rat
SD
M,
F
20
Males
0
0.57
6.1
49
224
Females
0
0.64
6.9
55
236
Drinking
water
90
days
Body
weight,
serum
chemistry,
histology
49
224
(
decreased
weight
gain,
mild
hepatic
lesions)
18
5.3
(
Liver
lesions
in
males)

Daniel
et
al.

(
1990)
Rat
SD
M,
F
10
0
50
100
200
Gavage
(
oil)
90
days
Body
weight,
clinical
signs,
serum
biochemistry,
gross
necropsy,
histology
­­
50
(
hepatic
and
renal
lesions)
20*
4.2*

(
kidney
cortex
degeneration
in
females)

NTP
(
1985)
Rat
F344/
N
M,
F
10
0
15
30
60
125
250
Gavage
(
oil)
13
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
histology
30
60
(
hepatic
lesions)
2.5
1.7
(
liver
fatty
metamorphosis
in
males)

NTP
(
1985)
Mouse
B6C3F
1
M,
F
10
0
15
30
60
125
250
Gavage
(
oil)
13
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
histology
125
250
(
renal
and
hepatic
lesions)
Not
modeled
­­

*
BMDL10
values
were
derived
using
duration­
adjusted
doses.

*
Modeled
using
Crump
benchmark
dose
software
Abbreviations:
SD,
Sprague­
Dawley
Draft
­
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or
Quote
February
20,
2003
VIII
­
42
significance
at
60
mg/
kg­
day,
and
this
dose
was
designated
the
LOAEL.
The
next
lower
dose
(
30
mg/
kg­
day)
was
designated
as
the
NOAEL.
BMD
analysis
using
the
BMDS
program
obtained
duration­
adjusted
BMD
and
BMDL
10
values
of
2.5
and
1.7
mg/
kg­
day,
respectively.
These
values
were
the
lowest
calculated
among
the
three
studies
for
which
BMD
analysis
was
conducted.

Using
the
NTP
(
1985)
rat
study,
the
Longer­
term
HA
for
the
10­
kg
child
is
calculated
as
follows:

Longer­
term
HA
=
(
1.7
mg/
kg­
day)(
10
kg)
=
0.17
mg/
L
(
rounded
to
0.2
mg/
L)
(
100)
(
1
L/
day)

where:

1.7
mg/
kg­
day
=
Duration­
adjusted
BMDL
10
based
on
hepatic
cell
vacuolization
in
rats
exposed
to
dibromochloromethane
by
oil
gavage
for
13
weeks
10
kg
=
Assumed
body
weight
of
a
child
100
=
Composite
uncertainty
factor
based
on
NAS/
OW
guidelines;
includes
a
factor
of
10
for
interspecies
variation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
The
Longer­
term
HA
for
a
70­
kg
adult
consuming
2
liters
of
water
per
day
is
calculated
according
to
the
following
equation:

Longer­
term
HA
=
(
1.7
mg/
kg­
day)(
70
kg)
=
0.60
mg/
L
(
rounded
to
0.6
mg/
L)
(
100)
(
2
L/
day)

For
purposes
of
comparison,
the
Longer­
term
HAs
were
also
calculated
using
the
conventional
NOAEL/
LOAEL
approach.
NTP
(
1985)
identified
a
NOAEL
of
30
mg/
kg­
day
by
based
on
the
absence
of
clinical
signs
or
histologic
alterations
in
rats
exposed
to
dibromochloromethane
by
corn
oil
gavage
for
13
weeks.
Using
a
duration
adjusted
dose
of
21
mg/
kg­
day
(
obtained
by
multiplying
the
nominal
dose
by
5/
7)
and
the
assumptions
for
body
weight
and
drinking
water
ingestion
described
above,
the
Longer­
term
HAs
would
be
2.1
and
7.5
mg/
kg­
day
for
a
10
kg
child
and
70
kg
adult,
respectively.
Draft
­
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or
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February
20,
2003
VIII
­
43
d.
Reference
Dose,
Drinking
Water
Equivalent
Level
and
Lifetime
Health
Advisory
This
section
reports
the
existing
RfD
value
for
dibromochloromethane
and
describes
the
derivation
of
the
RfD
for
this
compound.
This
section
also
describes
the
calculation
of
Drinking
Water
Equivalent
Level
and
Lifetime
Health
Advisory
values
which
require
the
RfD
as
input.
For
this
document,
new
and
existing
studies
were
reviewed
and
appropriate
candidate
data
were
selected
for
benchmark
dose
(
BMD)
modeling.
The
results
of
BMD
modeling
were
used
in
conjunction
with
appropriate
uncertainty
factors
to
calculate
the
RfD.
A
comparison
of
the
RfD
derived
using
the
BMD
approach
to
the
results
obtained
using
the
conventional
NOAEL/
LOAEL
approach
is
also
provided.

Description
of
the
Existing
RfD
The
existing
RfD
for
dibromochloromethane
is
0.02
mg/
kg­
day
(
IRIS,
1992).
This
value
was
derived
using
a
duration­
adjusted
NOAEL
of
21.4
mg/
kg­
day
identified
for
the
occurrence
of
hepatic
lesions
in
F344/
N
rats
administered
dibromochloromethane
by
corn
oil
gavage,
5
days/
week
for
13
weeks
.
An
uncertainty
factor
of
1000
was
used
to
account
for
extrapolation
from
animal
data,
for
protection
of
sensitive
human
subpopulations,
and
for
use
of
a
subchronic
study.

Identification
of
Candidate
Studies
for
Derivation
of
the
RfD
Candidate
studies
considered
for
derivation
of
the
RfD
for
dibromochloromethane
are
summarized
in
Table
VIII­
11
(
below).
Tobe
et
al.
(
1982)
administered
microencapsulated
dibromochloromethane
in
the
diet
to
Wistar
rats
for
24
months
at
dose
levels
that
ranged
from
12
to
278
mg/
kg­
day.
Although
the
study
identified
a
NOAEL
and
a
LOAEL
of
12
and
49
mg/
kgday
respectively,
based
on
decreased
body
weight,
changes
in
clinical
chemistry
parameters,
and
gross
liver
appearance
in
males,
a
histopathological
examination
was
not
conducted.
NTP
(
1985)
investigated
the
chronic
oral
toxicity
of
dibromochloromethane
in
F344/
N
rats
and
B6C3F
1
mice.
Only
LOAEL
values
were
identified
in
these
studies.
Specifically,
the
rat
study
identified
a
LOAEL
of
40
mg/
kg­
day
based
on
histologic
lesions
in
both
male
and
female
rats
(
e.
g.,
fatty
change),
and
the
mouse
study
identified
a
LOAEL
of
50
mg/
kg­
day
based
on
lesions
in
the
liver
(
fatty
metamorphosis)
and
the
thyroid
(
follicular
cell
hyperplasia)
in
the
female
mice.
A
thirteenweek
oral
exposure
study
in
rats
(
NTP,
1985)
examined
toxicity
at
a
wider
range
of
doses
than
the
chronic
studies
and
identified
NOAEL
and
LOAEL
values
of
30
and
60
mg/
kg­
day,
respectively,
for
histopathological
changes
in
the
liver.

Method
of
Analysis
Selected
data
from
the
candidate
studies
were
analyzed
using
the
benchmark
dose
(
BMD)
modeling
approach.
Initially,
data
sets
for
potentially
sensitive
endpoints
were
selected
as
described
in
U.
S.
EPA
(
1998b)
and
analyzed
using
the
Crump
Benchmark
Dose
Modeling
Software
(
K.
S.
Crump,
Inc.).
Results
of
this
preliminary
analysis
are
summarized
in
Table
VIII­
12.
Following
the
release
of
Version
1.2
of
the
BMDS
program
(
U.
S.
EPA,
2000a),
a
subset
of
the
most
sensitive
endpoints
identified
using
the
Crump
software
was
reanalyzed
in
accordance
with
proposed
U.
S.
EPA
(
2000b)
recommendations.
An
advantage
of
analysis
with
the
BMDS
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44
software
is
that
several
additional
models
are
available
to
fit
the
data.
The
results
of
the
analysis
using
the
BMDS
software
are
included
in
Table
VIII­
11.

Choice
of
Principal
Study
and
Critical
Effect
for
the
RfD
Two
studies
were
identified
as
strong
candidates
for
selection
as
the
principal
study.
The
NTP
(
1985)
subchronic
study
evaluated
toxicological
effects
in
male
and
female
rats
at
five
concentrations
of
dibromochloromethane
(
15,
30,
60,
125,
and
250
mg/
kg­
day)
in
addition
to
the
control.
The
chemical
was
administered
by
gavage
in
oil
on
five
days
per
week
for
13
weeks.
A
relatively
small
sample
size
of
10
animals/
treatment
group
was
utilized.
The
endpoints
evaluated
included
clinical
signs,
body
weight,
serum
biochemistry,
and
histopathological
changes
in
organs.
NOAEL
and
LOAEL
values
of
30
and
60
mg/
kg­
day,
respectively,
were
identified
on
the
basis
of
hepatic
lesions
using
the
conventional
approach.
Analysis
of
the
data
for
fatty
metamorphosis
in
the
liver
using
the
BMDS
program
resulted
in
a
duration­
adjusted
BMD
of
2.7
mg/
kg­
day
for
this
endpoint,
with
a
corresponding
duration­
adjusted
BMDL
10
of
1.7
mg/
kg­
day.
A
strength
of
this
study
with
respect
to
BMD
modeling
was
the
use
of
additional
doses
at
the
lower
end
of
the
dose­
response
range.
Inclusion
of
these
doses
permits
more
accurate
characterization
of
the
shape
of
the
dose­
response
curve
and
thus
less
uncertainty
in
the
range
of
interest.

The
second
candidate
for
selection
as
the
principle
study
for
derivation
of
the
RfD
was
the
chronic
study
conducted
by
NTP
(
1985).
This
study
evaluated
dibromochloromethane
effects
at
administered
doses
of
0,
40
and
80
mg/
kg­
day.
The
chemical
was
administered
by
gavage
in
oil
on
five
days
per
week
for
104
weeks.
The
endpoints
evaluated
included
clinical
signs,
body
weight,
serum
biochemistry,
and
histopathological
changes
in
organs,
and
a
LOAEL
of
40
mg/
kgday
based
on
hepatic
lesions
was
identified
using
the
conventional
approach.
Analysis
of
the
data
for
fatty
metamorphosis
in
the
liver
using
the
BMDS
program
resulted
in
a
duration­
adjusted
BMD
of
2.5
mg/
kg­
day
for
this
endpoint,
with
a
corresponding
duration­
adjusted
BMDL
10
of
1.6
mg/
kg­
day
based
on
duration
adjusted
doses.

A
potential
weakness
of
the
NTP
(
1985)
chronic
study
is
the
lack
of
dose­
response
information
at
administered
doses
less
than
40
mg/
kg­
day.
A
priori,
the
lack
of
information
regarding
the
shape
of
the
curve
at
low
doses
would
be
expected
to
result
in
greater
uncertainty
(
and
thus
wider
confidence
limits)
in
the
estimate
of
the
chronic
BMD.
However,
the
BMD
and
BMDL
10
values
calculated
for
fatty
metamorphosis
in
the
subchronic
and
chronic
studies
are
closely
similar.
This
observation
suggests
that
there
is
little
potential
for
cumulative
effects
on
the
occurrence
of
this
lesion.
The
slight
differences
in
the
values
may
reflect
both
experimental
uncertainty
and
uncertainty
in
modeling.

Both
studies
were
considered
appropriate
for
derivation
of
the
RfD.
The
NTP
(
1985)
investigation
of
chronic
toxicity
in
rats
was
selected
as
the
principle
study
on
the
basis
of
its
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45
Table
VIII­
11
Summary
of
Candidate
Studies
for
Derivation
of
the
RfD
for
Dibromochloromethane
Reference
Species
Sex
n
Dose
(
mg/
kg­
day)
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
*

Tobe
et
al.

(
1982)
Rat
Wistar
M,
F
40
Males
0
12
49
196
Females
0
17
70
278
Diet
24
months
Body
weight,
serum
biochemistry,
gross
pathology
12
49
(
serum
enzyme
changes
and
altered
liver
appearance)
Not
modeled
­­

NTP
(
1985)
Rat
F344/
N
M,
F
10
0
15
30
60
125
250
Gavage
(
oil)
13
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
serum
biochemistry,
gross
necropsy,
histology
30
60
(
hepatic
lesions)
2.5
1.7
(
fatty
metamorphosis
in
liver
of
males)

NTP
(
1985)
Rat
F344/
N
M,
F
50
0
40
80
Gavage
(
oil)
104
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
gross
necropsy,

histology
­­
40
(
hepatic
lesions)
2.7
1.6
(
fatty
changes
in
liver
of
males)

NTP
(
1985)
Mouse
B6C3F
1
M,
F
50
0
50
100
Gavage
(
oil)
105
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
gross
necropsy,

histology
­­
50
(
hepatic
lesions)
9.1*
7.1*

(
thyroid
follicular
cell
hyperplasia
in
females)

Borzelleca
and
Carchman
(
1982)
**
Mouse
ICR
Swiss
­
10M
30F
0
17
171
685
Drinking
water
27
weeks
Maternal
body
weight,
gross
pathology,
fetal
weight,
survival,

teratogenicity
17
(
marginal)
171
(
maternal
toxicity,

possible
fetotoxicity)
Not
modeled
(
insuff.
data
provided
in
publication)
­­
Table
VIII­
11
(
cont.)

Reference
Species
Sex
n
Dose
(
mg/
kg­
day)
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
*

Draft
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46
NTP
(
1996)

**
Rat
SD
M,
F
10
males
0
4.2
12.4
28.2
Group
A
females
0
6.3
17.4
46.0
Group
B
females
0
7.1
20.0
47.8
Drinking
water
29
days
Body
weight,
serum
chemistry,
hematology,
gross
necropsy,
histology,

sperm
evaluation
28
­­
Not
modeled
­­

*
BMDL10
values
were
derived
using
duration­
adjusted
doses.

**
These
studies
have
been
included
in
this
table
because
they
are
reproductive/
developmental
studies
and
would
be
considered
relevant
for
derivation
of
the
RfD.
However,
Borzelleca
and
Carchman
(
1982)
found
only
marginal
evidence
for
developmental
toxicity
at
the
low
dose
level
and
the
NTP
(
1996)

study
did
not
observe
any
reproductive
or
developmental
effects
at
the
dose
levels
evaluated.

*
Modeled
using
Crump
BMD
software
Abbreviations:
SD,
Sprague­
Dawley
Draft
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47
Table
VIII­
12
Results
of
Preliminary
BMD
Modeling
of
Selected
Data
from
NTP
(
1985)
Studies
Study
Endpoint
Modeled
BMDL10
(
mg/
kg­
day)
*

Subchronic
NTP
(
1985)
rat
study
Fatty
metamorphosis
in
liver
of
male
rats
0.93
Chronic
NTP
(
1985)
rat
study
Fatty
metamorphosis
in
liver
of
male
rats
1.16
Fatty
metamorphosis
in
liver
of
female
rats
No
acceptable
fit
"
Ground
glass"
cytoplasm
in
liver
of
male
rats
4.93
Nephrosis
in
liver
of
female
rats
17
Chronic
NTP
(
1985)
mouse
study
Fatty
metamorphosis
in
liver
of
female
mice
7.68
Thyroid
follicular
cell
hyperplasia
in
female
mice
7.09
*
BMD
modeling
was
conducted
on
duration­
adjusted
values
using
the
Crump
BMD
software.

longer
duration.
The
critical
endpoint
is
hepatotoxicity,
as
evidenced
by
the
occurrence
of
fatty
metamorphosis
in
the
livers
of
dibromochloromethane­
treated
animals.
This
effect
was
strongly
dose­
dependent,
with
incidences
of
27/
50,
47/
50,
and
49/
50
at
the
duration­
adjusted
doses
of
0,
29,
and
57
mg/
kg­
day.
Selection
of
this
study
is
strongly
supported
by
the
similar
BMD
calculated
for
the
same
effect
in
the
NTP
(
1985)
subchronic
study.

Derivation
of
the
RfD
The
duration­
adjusted
BMDL
10
value
from
the
chronic
NTP
(
1985)
rat
study
was
selected
as
the
most
appropriate
basis
for
derivation
of
the
RfD
for
dibromochloromethane.
The
RfD
is
calculated
using
the
following
equation:

RfD
=
(
1.6
mg/
kg­
day)
=
0.016
mg/
kg­
day
(
rounded
to
0.02
mg/
kg­
day)
(
100)

where:

1.6
mg/
kg­
day
=
Duration­
adjusted
BMDL
10
based
on
fatty
changes
in
the
liver
of
male
rats
100
=
Composite
uncertainty
factor
based
on
NAS/
OW
guidelines;
includes
a
factor
of
10
interspecies
extrapolation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
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­
48
A
composite
UF
of
100
was
used.
The
standard
factors
of
10
were
used
for
interspecies
extrapolation
and
for
protection
of
sensitive
subpopulations.
Furthermore,
no
additional
uncertainty
factor
was
needed
to
account
for
an
incomplete
database.
The
database
for
dibromochloromethane
includes
a
two­
generation
study
in
ICR
Swiss
mice
(
Borzelleca
and
Carchman
1982),
a
developmental
toxicity
study
in
Sprague­
Dawley
rats
(
Ruddick
et
al.,
1983),
and
a
short­
term
reproductive
and
developmental
toxicity
study
in
rats
(
NTP,
1996).
Therefore,
the
database
is
considered
nearly
complete
despite
the
lack
of
a
developmental
toxicity
study
in
a
second
species.

The
DWEL
for
dibromochloromethane
is
calculated
as
follows:

DWEL
=
(
0.02
mg/
kg­
day)
(
70
kg)
=
0.7
mg/
L
(
700
µ
g/
L)
2
L/
day
where:

0.02
mg/
kg­
day
=
RfD
70
kg
=
Assumed
weight
of
an
adult
2
L/
day
=
Assumed
water
consumption
by
a
70­
kg
adult
Lifetime
Health
Advisory
The
Lifetime
Health
Advisory
(
HA)
represents
that
portion
of
an
individual's
total
exposure
that
is
attributed
to
drinking
water
and
is
considered
protective
of
noncarcinogenic
health
effects
over
a
lifetime
of
exposure.
Dibromochloromethane
is
classified
with
respect
to
carcinogenic
potential
as
Group
C:
Possible
human
carcinogen.
The
Lifetime
Health
Advisory
(
HA)
is
therefore
calculated
as
follows:

Lifetime
HA
=
(
0.7
mg/
L)
(
0.8)
=
0.06
mg/
L
(
60
µ
g/
L)
10
where:

0.7
mg/
kg­
day
=
DWEL
0.8
=
Relative
Source
Contribution
(
RSC),
the
proportion
of
the
total
daily
exposure
contributed
by
the
dibromochloromethane
in
drinking
water
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VIII
­
49
10
=
Uncertainty
factor
used
in
accordance
with
U.
S.
EPA
policy
for
Group
C
contaminants
to
account
for
possible
carcinogenicity
Alternative
Approach
for
Derivation
of
the
RfD
An
alternative
approach
to
the
derivation
of
the
RfD
is
use
of
the
conventional
NOAEL/
LOAEL
method.
The
subchronic
oral
exposure
study
conducted
by
NTP
(
1985)
identified
a
NOAEL
of
30
mg/
kg­
day.
Using
this
value,
a
duration
adjustment
factor
of
5/
7,
and
a
composite
uncertainty
factor
of
1000
(
includes
factors
of
10
for
interspecies
extrapolation,
protection
of
sensitive
subpopulations,
and
use
of
a
subchronic
study),
the
resulting
RfD
is
0.02
mg/
kg­
day
(
the
same
value
as
derived
using
the
BMD
approach).
The
corresponding
DWEL
is
0.7
mg/
L,
assuming
an
adult
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L/
day.

2.
Carcinogenic
Effects
a.
Categorization
of
Carcinogenic
Potential
Previous
Evaluations
of
Carcinogenic
Potential
The
Carcinogenic
Risk
Assessment
Verification
Endeavor
(
CRAVE)
group
of
the
U.
S.
EPA
reviewed
the
available
evidence
on
the
carcinogenicity
of
the
brominated
trihalomethanes
and
assigned
dibromochloromethane
to
Group
C:
possible
human
carcinogen
(
IRIS,
1992).
This
classification
reflects
inadequate
human
data
and
limited
evidence
of
carcinogenicity
in
animals.

Based
on
the
1996
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
published
in
1996
(
U.
S.
EPA,
1996),
dibromochloromethane
is
classified
as
cannot
be
determined.
This
descriptor
is
considered
appropriate
when
there
are
no
or
inadequate
data
in
humans,
and
limited
evidence
for
carcinogenicity
in
animals.

IARC
(
1999c)
has
recently
re­
evaluated
the
carcinogenic
potential
of
dibromochloromethane.
IARC
concluded
that
there
is
limited
evidence
of
carcinogenicity
in
experimental
animals
and
inadequate
evidence
in
humans
for
dibromochloromethane.
Dibromochloromethane
is
therefore
classified
as
Group
3:
not
classifiable
as
to
carcinogenicity
in
humans.

Categorization
of
Carcinogenic
Potential
Under
the
Proposed
1999
Cancer
Guidelines
Cancer
Hazard
Summary
Under
the
proposed
guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
dibromochloromethane
shows
suggestive
evidence
of
carcinogenicity,
but
not
sufficient
to
assess
human
carcinogenic
potential.
This
descriptor
is
appropriate
when
the
evidence
from
human
or
Draft
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February
20,
2003
VIII
­
50
animal
data
is
suggestive
of
carcinogenicity,
which
raises
a
concern
for
carcinogenic
effects
but
is
not
judged
sufficient
for
a
conclusion
as
to
human
carcinogenic
potential.
This
finding
is
based
on
the
weight
of
experimental
evidence
in
animal
models
which
indicate
limited
or
equivocal
evidence
of
carcinogenicity.
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2003
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­
51
Supporting
Information
for
Cancer
Hazard
Assessment
Human
Data
The
information
on
the
carcinogenicity
of
dibromochloromethane
from
human
studies
is
inadequate.
There
are
no
epidemiological
data
specifically
relating
increased
incidence
of
cancer
to
exposure
to
dibromochloromethane.
There
are
equivocal
epidemiological
data
describing
a
weak
association
of
chlorinated
drinking
water
exposures
with
increased
incidences
of
bladder,
rectal,
and
colon
cancer.
U.
S.
EPA
has
determined
that
these
studies
cannot
attribute
the
observed
effects
to
a
single
compound,
as
chlorinated
water
contains
numerous
other
disinfection
byproducts
that
are
potentially
carcinogenic.

Animal
Data
The
carcinogenicity
of
dibromochloromethane
in
male
and
female
animals
has
been
investigated
in
a
well­
designed
and
conducted
corn
oil
gavage
study
conducted
in
rats
and
mice,
a
dietary
exposure
study
in
rats,
a
drinking
water
study
in
mice,
and
a
feed
study
in
rats.
No
data
are
available
on
the
carcinogenic
potential
of
dibromochloromethane
administered
via
the
inhalation
or
dermal
routes.

In
the
corn
oil
gavage
study
(
NTP,
1985),
the
incidence
of
hepatocellular
adenomas
and
carcinomas
and
combined
adenomas
and
carcinomas
was
significantly
increased
in
high­
dose
female
mice
and
the
incidence
of
hepatocellular
adenomas
was
significantly
increased
in
high
dose
male
mice.
No
evidence
was
observed
for
carcinogenicity
in
male
or
female
rats
under
the
experimental
conditions
employed.
Voronin
(
1987)
did
not
observe
significant
increases
in
mice
treated
with
dibromochloromethane
in
drinking
water
for
104
weeks.
Tobe
et
al.
(
1982)
reported
no
increase
in
gross
tumors
in
rats
treated
exposed
to
dibromochloromethane
in
the
diet
for
two
years.

Structural
Analogue
Data
Dibromochloromethane
is
structurally
related
to
trihalomethanes
that
have
shown
varying
degrees
of
carcinogenic
potential
in
rodents.
Chloroform,
the
most
extensively
characterized
trihalomethane,
is
reported
to
be
carcinogenic
at
high
doses
in
several
chronic
animal
bioassays,
with
significant
increases
in
the
incidence
of
liver
tumors
in
male
and
female
mice
and
significant
increases
in
the
incidence
of
kidney
tumors
in
male
rats
and
mice
(
U.
S.
EPA,
2001).
The
occurrence
of
tumors
in
animals
exposed
to
chloroform
is
demonstrably
species­,
strain­,
and
gender­
specific,
and
has
only
been
observed
under
dose
conditions
that
caused
cytotoxicity
and
regenerative
cell
proliferation
in
the
target
organ.
The
cancer
database
for
structurally­
related
brominated
trihalomethanes
is
more
limited,
but
includes
well­
conducted
studies
performed
by
the
National
Toxicology
Program.
In
a
two­
year
corn
oil
gavage
study
of
bromoform,
NTP
(
1989a)
found
clear
evidence
for
carcinogenicity
in
female
rats
and
some
evidence
of
carcinogenicity
based
on
occurrence
of
tumors
of
the
large
intestine.
In
a
two­
year
corn
oil
gavage
study
of
bromodichloromethane,
NTP
(
1987)
found
clear
evidence
of
carcinogenicity
in
male
and
female
rats
(
tumors
of
the
large
intestine),
male
mice
(
kidney
tumors),
and
female
mice
(
liver
tumors).
In
other
bioassays,
George
et
al.
(
2002)
observed
a
significantly
increased
prevalence
of
neoplastic
Draft
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20,
2003
VIII
­
52
lesions
in
the
liver
of
male
rats
at
the
lowest
dose
of
bromodichloromethane
administered
in
drinking
water,
but
not
at
higher
doses.
Tumasonis
et
al.
(
1985)
reported
significantly
increased
incidences
of
hepatic
neoplastic
nodules,
hepatic
adenofibrosis,
and
lymphosarcoma
in
female
rats
exposed
to
bromodichloromethane
in
drinking
water.

Other
Key
Data
Dibromochloromethane
is
formed
as
a
byproduct
of
drinking
water
disinfection
with
chlorine.
Exposure
to
dibromochloromethane
may
occur
via
ingestion
of
tap
water,
via
dermal
contact
during
showering
or
bathing,
or
by
inhalation
of
dibromochloromethane
volatilized
during
household
activities.
Absorption
of
single
oral
doses
appears
to
be
extensive.
Dibromochloromethane
is
rapidly
metabolized
and
eliminated
predominately
as
expired
volatiles,
CO
2,
or
CO.
Only
a
small
amount
(
less
than
10%)
is
eliminated
in
urine
or
in
feces.
No
comprehensive
tissue
data
are
available
regarding
the
bioaccumulation
or
retention
of
dibromochloromethane
following
repeated
exposure.
However,
because
of
the
rapid
metabolism
and
excretion
of
dibromochloromethane,
marked
accumulation
and
retention
is
not
expected.

Dibromochloromethane
itself
is
not
directly
reactive
with
DNA.
Metabolism
to
reactive
species
is
a
prerequisite
for
toxicity,
as
inferred
from
metabolic
induction
and
inhibition
studies.
In
vitro
and
in
vivo
studies
of
the
mutagenic
and
genotoxic
potential
of
bromodichloromethane
have
yielded
both
positive
and
negative
results.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid)
and,
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
Study
results
for
the
mutagenicity
of
dibromochloromethane
are
mixed,
and
the
overall
evidence
for
mutagenicity
of
this
chemical
is
judged
to
be
inconclusive
(
U.
S.
EPA,
1994b).
Recent
studies
conducted
with
strains
of
Salmonella
engineered
to
express
rat
theta­
class
glutathione
Stransferase
suggest
that
mutagenicity
of
the
brominated
trihalomethanes
may
be
mediated
by
glutathione
conjugation.

Mode
of
Action
Limited
or
equivocal
evidence
has
been
obtained
for
the
carcinogenic
potential
of
dibromochloromethane.
Data
to
support
primary
mode
of
action
for
tumor
development
in
the
liver
of
mice
exposed
to
dibromochloromethane
are
lacking.
In
the
absence
of
such
information,
combined
with
an
inconclusive
weight­
of­
evidence
evaluation
for
genotoxicity,
the
mode
of
action
for
tumor
development
is
assumed
to
be
a
linear
process.
The
processes
leading
to
tumor
formation
in
animals
are
expected
to
be
relevant
to
humans.

Conclusion
Under
the
proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
dibromochloromethane
shows
suggestive
evidence
of
carcinogenicity,
but
not
sufficient
to
assess
human
carcinogenic
potential
by
the
oral
route.
This
weight­
of­
evidence
evaluation
is
based
on
1)
limited
or
equivocal
evidence
of
carcinogenicity
in
mice,
but
not
rats,
treated
by
oral
pathways;
2)
lack
of
epidemiological
data
specific
to
dibromochloromethane
and
equivocal
data
for
drinking
Draft
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20,
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VIII
­
53
water
drinking
water
exposures
that
cannot
reliably
be
attributed
to
dibromochloromethane
among
multiple
other
disinfection
byproducts;
3)
inconclusive
results
for
many
of
the
available
genotoxicity
and
mutagenicity
tests;
and
4)
metabolism
and
mode
of
action
that
are
reasonably
expected
to
be
similar
to
those
of
structurally
related
compounds
that
induce
tumors
in
experimental
animals.
Although
no
cancer
data
exist
for
exposures
via
the
dermal
or
inhalation
pathways,
the
weight­
of­
evidence
conclusion
is
considered
to
be
applicable
to
these
pathways
as
well.
The
finding
for
inhalation
is
based
on
the
observation
that
patterns
of
metabolizing
enzyme
activity
in
male
rats
for
the
related
trihalomethane
bromodichloromethane
are
similar
for
exposure
via
the
inhalation
and
gavage
routes.
Dibromochloromethane
absorbed
through
the
skin
is
expected
to
be
metabolized
and
cause
toxicity
in
much
the
same
way
as
dibromochloromethane
absorbed
by
the
oral
and
inhalation
routes.

b.
Choice
of
Study
for
Quantification
of
Carcinogenic
Risk
The
proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
do
not
indicate
dose­
response
assessment
for
chemicals
for
which
there
is
suggestive
evidence
of
carcinogenicity,
but
not
sufficient
to
assess
human
carcinogenic
potential.
However,
the
single
oral
exposure
study
with
positive
tumor
data
for
dibromochloromethane
suggests
significant
cancer
potency
for
this
compound
in
mice.
A
quantitative
assessment
of
potency
was
therefore
considered
appropriate.

In
the
absence
of
other
carcinogenicity
data,
hepatic
tumor
incidence
in
female
mice
was
selected
for
estimation
of
carcinogenic
risks
associated
with
dibromochloromethane.
These
data
were
obtained
in
an
NTP
(
1985)
study
in
which
dibromochloromethane
was
administered
in
corn
oil
to
male
and
female
B6C3F
1
mice
(
50
mice/
sex/
dose)
by
gavage
5
times/
week
for
104
to
105
weeks.
The
administered
doses
were
0,
50,
or
100
mg/
kg­
day.
Survival
of
dosed
female
mice
was
comparable
to
that
of
the
corresponding
vehicle­
control
groups.
High­
dose
male
mice
had
lower
survival
rates
than
the
vehicle
controls.
At
week
82,
nine
high­
dose
male
mice
died
of
an
unknown
cause.
An
inadvertent
overdose
of
dibromochloromethane
given
to
low­
dose
male
and
female
mice
at
week
58
killed
35
male
mice,
but
apparently
did
not
affect
the
female
mice.
The
low­
dose
male
mouse
group
was,
therefore,
considered
to
be
unsuitable
for
analysis
of
neoplasms.
Compound­
related
nonneoplastic
lesions
were
found
in
primarily
in
the
livers
of
males
(
hepatocytomegaly,
necrosis,
fatty
metamorphosis)
and
females
(
calcification
and
fatty
metamorphosis).
Nephrosis
was
also
observed
in
male
mice.
Statistically
significant
increases
in
the
incidence
of
hepatocellular
adenomas
and
in
the
combined
incidence
of
adenomas
and
carcinomas
were
observed
in
high­
dose
female
mice.
In
male
mice,
a
statistically
significant
increase
in
the
incidence
of
hepatocellular
carcinomas
and
combined
adenomas
and
carcinomas
was
observed
in
the
high­
dose
group;
however,
due
to
the
overdose
of
dibromochloromethane
in
the
mid­
dose
group,
the
authors
considered
the
tumor
incidence
data
inadequate
for
tumor
analysis.
Tumor
incidence
data
from
this
study
are
presented
in
Table
VIII­
13.

c.
Extrapolation
model
The
LMS
model
(
U.
S.
EPA,
1986)
and
the
default
linear
approach
described
by
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1996;
1999)
were
used
to
quantify
the
risk
associated
with
exposure
to
dibromochloromethane.
Data
for
the
mutagenicity
and
Draft
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February
20,
2003
VIII
­
54
genotoxicity
of
dibromochloromethane
are
mixed
(
see
Section
V.
F.
2).
U.
S.
EPA
(
1994b)
has
previously
determined
that
the
weight
of
evidence
for
dibromochloromethane
mutagenicity
and
genotoxicity
is
inconclusive.
At
the
present
time
there
is
insufficient
evidence
to
establish
with
certainty
that
dibromochloromethane
exerts
its
carcinogenic
effects
via
a
non­
genotoxic
mechanism.
Thus,
use
of
linear
approaches
was
considered
appropriate
for
quantification
of
cancer
risk
associated
with
exposure
to
this
compound.

Table
VIII­
13
Frequencies
of
Liver
Tumors
in
Mice
Administered
Dibromochloromethane
in
Corn
Oil
for
105
Weeks
­
Adapted
from
NTP
(
1985)

Treatment
(
mg/
kg­
day)
Sex
Adenoma
Carcinoma
Adenoma
or
Carcinoma
(
combined)

Vehicle
Control
M
F
14/
50
2/
50
10/
50
4/
50
23/
50
6/
50
50
M
F
­­
a
4/
49
­­
6/
49
­­
10/
49
100
M
F
10/
50
11/
50b
19/
50b
8/
50
27/
50c
19/
50d
a
Male
low­
dose
group
was
inadequate
for
statistical
analysis.
b
p
<
0.05
relative
to
controls.
c
p
<
0.01
(
life
table
analysis);
p
=
0.065
(
incidental
tumor
test)
relative
to
controls.
d
p
<
0.01
relative
to
controls.

d.
Cancer
Potency
and
Unit
Risk
The
only
tumor
data
available
for
dibromochloromethane
are
for
liver
tumors
in
female
B6C3F
1
mice
(
NTP,
1985).
NAS
(
1987)
previously
utilized
the
tumor
frequency
data
reported
by
NTP
(
1985)
to
calculate
an
excess
lifetime
cancer
unit
risk
of
8.3
x
10­
7.
The
linearized
multistage
model
was
utilized,
with
the
assumption
that
1
L
of
water
per
day
containing
1
µ
g/
L
of
dibromochloromethane
was
ingested.
Based
on
this
calculation,
the
concentration
associated
with
a
risk
of
10­
6
is
0.6
µ
g/
L,
assuming
consumption
of
2
L
of
water
per
day.

Other
available
estimates
of
cancer
risks
are
summarized
in
Table
VIII­
14.
U.
S.
EPA
(
1994b)
reported
a
slope
factor
of
8.4
x
10­
2
(
mg/
kg­
day)­
1
calculated
from
the
NTP
(
1985)
data
in
the
absence
of
other
appropriate
tumorigenicity
data
for
dibromochloromethane
(
IRIS,
1992).
This
value
was
derived
using
the
LMS
model
(
extra
risk)
and
a
scaling
factor
of
body
weight2/
3,
as
specified
in
the
1986
Guidelines
for
Carcinogenic
Risk
Assessment
(
U.
S.
EPA,
1986).
The
reported
unit
risk
and
10­
5
risk
concentration
were
2.4
x
10­
6
(
µ
g/
L)­
1
and
4
µ
g/
L,
respectively.

A
slope
factor
of
4.3
x
10­
2
(
mg/
kg­
day)­
1
(
U.
S.
EPA,
1998b)
was
derived
using
the
LMS
model
and
a
scaling
factor
of
body
weight3/
4.
The
use
of
body
weight3/
4
as
the
scaling
factor
is
consistent
with
recommendations
in
(
U.
S.
EPA.,
1992b).
A
unit
risk
value
of
1.2
×
10­
6
(
µ
g/
L)­
1
was
estimated
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
Draft
­
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or
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February
20,
2003
VIII
­
55
This
estimate
was
used
to
calculate
a
drinking
water
10­
5
risk
concentration
of
8
µ
g/
L
(
0.8
µ
g/
L
at
10­
6
risk).

Table
VIII­
14
Carcinogenic
Risk
Estimates
for
Dibromochloromethane
Method
of
Estimation
Tumor
Site
Species
Sex
Slope
Factor
(
mg/
kg­
day)­
1
Unit
Risk
(
µ
g/
L)­
1
10­
5
Risk
Conc.
(
µ
g/
L)
LED
10
(
µ
g/
kg­
day)

LMS
Method
Using
BW3/
4
Conversion
U.
S.
EPA
(
1998b)
Liver
Mouse
F
4.3
×
10­
2
1.2
×
10­
6
8
­

LMS
Method
Using
BW2/
3
Conversion
U.
S.
EPA
(
1994b)*
Liver
Mouse
F
8.4
×
10­
2
2.4
×
10­
6
4
­

LED
10/
Linear
Method
U.
S.
EPA
(
1998b)
Liver
Mouse
F
4.0
×
10­
2
1.2x10­
6
9
2.5
×
10
3
*
Adapted
from
IRIS
(
1992)

Cancer
risk
estimates
were
also
obtained
using
the
LED
10
(
the
lower
95%
confidence
limit
on
a
dose
associated
with
10%
extra
risk)
for
hepatic
tumors
and
assuming
a
linear
mode
of
action
for
the
carcinogenicity
of
dibromochloromethane
(
Table
VIII­
14).
A
cancer
potency
value
of
4.0
x
10­
2
(
mg/
kg­
day)­
1
was
derived
using
this
approach.
A
unit
risk
of
1.2
x
10­
6
(
µ
g/
L)­
1
was
calculated
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
9
µ
g/
L
associated
with
a
10­
5
risk
(
0.9
µ
g/
L
for
10­
6
risk).
These
values
are
similar
to
values
derived
using
the
LMS
approach
with
body
weight
scaling
to
the
3/
4
power.

The
use
of
a
corn
oil
vehicle
in
the
NTP
(
1985)
study
from
which
these
data
are
derived
contributes
uncertainty
regarding
the
relevance
of
this
value
to
exposure
via
drinking
water.
The
U.
S.
EPA
plans
to
seek
data
on
the
tumorigenicity
of
dibromochloromethane
in
water
in
order
to
clarify
this
issue.
Draft
­
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or
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February
20,
2003
VIII
­
56
C.
Bromoform
1.
Noncarcinogenic
effects
a.
One­
day
Health
Advisory
Acute
toxicity
information
on
bromoform
is
limited
and
no
data
suitable
for
BMD
modeling
were
identified.
Some
information
is
available
on
the
former
medicinal
use
of
bromoform
in
humans.
In
the
past,
oral
doses
of
bromoform
were
used
as
a
sedative
for
children
with
whooping
cough.
Doses
were
typically
one
drop
(
approximately
180
mg)
given
three
to
six
times
per
day
(
Burton­
Fanning,
1901).
This
treatment
usually
resulted
in
mild
sedation
in
children,
although
a
few
rare
cases
of
death
or
near­
death
(
believed
to
be
due
to
accidental
overdoses)
have
been
reported
(
e.
g.,
Dwelle,
1903;
Benson,
1907).
Based
on
a
dose
of
540
mg/
day
given
to
a
10­
kg
child,
the
LOAEL
for
mild
sedation
is
about
54
mg/
kg­
day.
Accordingly,
the
one
day­
HA
for
bromoform
is
calculated
according
to
the
following
equation:

One­
day
HA
=
(
54
mg/
kg­
day)(
10
kg)
=
5.4
mg/
L
(
rounded
to
5
mg/
L)
(
100)
(
1
L/
day)

where:

54
mg/
kg­
day
=
LOAEL
based
on
sedation
in
children
given
oral
doses
of
bromoform
10
kg
=
Assumed
weight
of
a
child
100
=
Composite
uncertainty
factor
based
on
NAS/
OW
guidelines.
Includes
a
factor
of
10
for
interspecies
variation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
b.
Ten­
day
health
Advisory
Candidate
studies
considered
for
derivation
of
the
Ten­
day
HA
are
summarized
in
Table
VIII­
15
(
below).
Condie
et
al.
(
1983)
administered
bromoform
by
gavage
to
male
CD­
1
mice
at
doses
ranging
from
72
to
289
mg/
kg­
day
for
14
days
and
identified
a
NOAEL
of
145
mg/
kg­
day
and
a
LOAEL
of
289
mg/
kg­
day.
The
LOAEL
is
based
on
changes
in
clinical
chemistry
and
on
minimal
to
moderate
histologic
changes
in
the
kidney
(
intratubular
mineralization,
epithelial
hyperplasia,
and
mesangial
hypertrophy
and
nephrosis)
and
in
the
liver
(
centrilobular
pallor,
mitotic
figures,
focal
inflammation,
and
cytoplasmic
vacuolization).
BMD
modeling
of
data
for
renal
mesangial
hypertrophy
calculated
BMD
and
BMDL
10
values
of
73
and
34
mg/
kg­
day,
respectively.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
57
Melnick
et
al.
(
1998)
administered
bromoform
to
female
B6C3F
1
mice
by
gavage
5
days/
week
for
3
weeks
and
identified
a
NOAEL
of
200
mg/
kg­
day
(
the
lowest
dose
tested)
and
a
LOAEL
of
500
mg/
kg­
day
based
on
histologic
changes
in
the
liver
(
hepatocyte
hydropic
degeneration).
The
duration­
adjusted
BMD
and
BMDL
10
values
for
this
endpoint
were
146
and
104
mg/
kg­
day,
respectively.

Munson
et
al.
(
1982)
identified
NOAEL
and
LOAEL
values
of
125
and
250
mg/
kg­
day,
respectively,
based
on
elevated
serum
enzyme
activity
in
mice.
BMD
modeling
was
not
conducted
for
this
endpoint,
since
it
was
not
considered
a
reliable
basis
for
the
Ten­
day
HA
in
the
absence
of
histopathological
data.
NTP
(
1989a)
identified
NOAEL
and
LOAEL
values
of
200
and
400
mg/
kg­
day,
respectively,
based
on
the
occurrence
of
stomach
nodules
in
rats
and
mice.
A
BMD
of
167
mg/
kg­
day
was
calculated
for
this
endpoint
in
mice,
with
a
corresponding
BMDL
10
of
66
mg/
kg­
day.
However,
occurrence
of
these
nodules
may
represent
a
portal
of
entry
effect.
Chu
et
al.
(
1982a)
identified
a
freestanding
NOAEL
of
80
mg/
kg­
day
in
a
drinking
water
study
conducted
in
rats.
Coffin
et
al.
(
2000)
identified
a
LOAEL
of
200
mg/
kg­
day
based
on
the
occurrence
of
liver
histopathology
and
increased
labeling
index.
The
data
of
Coffin
et
al.
were
not
modeled
because
other
studies
used
lower
doses
and
were
thus
able
to
better
characterize
the
low­
dose
portion
of
the
dose
response
curve.

The
study
conducted
by
Aida
et
al.
(
1992a)
assessed
toxicity
in
Wistar
rats
administered
bromoform
microencapsulated
in
the
diet
at
doses
ranging
from
56
to
728
mg/
kg­
day.
The
duration
of
the
study
was
one
month.
This
study
identified
a
NOAEL
of
56
mg/
kg­
day
and
a
LOAEL
of
208
mg/
kg­
day
based
on
clinical
chemistry
changes
and
histologic
changes
in
the
liver
(
cell
vacuolization
and
swelling)
of
females.
BMD
modeling
of
results
for
liver
cell
vacuolization
in
female
rats
calculated
BMD
and
BMDL
10
values
of
16
and
2.3
mg/
kg­
day,
respectively.
These
were
the
lowest
values
observed
among
modeling
results
for
candidate
studies
for
the
10­
day
HA.
On
this
basis,
and
because
histopathological
changes
in
the
liver
are
considered
a
sensitive
indicator
of
brominated
trihalomethane
toxicity,
the
study
conducted
by
Aida
et
al.
(
1992a)
was
considered
the
best
choice
for
derivation
of
the
Ten­
day
HA.
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
58
Table
VIII­
15
Summary
of
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Bromoform
Reference
Species
Sex
n
Route
Dose
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Aida
et
al.

(
1992a)
Rat
Wistar
M,
F
7
Feed
Males
0
62
187
618
Females
0
56
208
728
1
month
Body
weight,

clinical
signs,

serum
biochemistry,

hematology,

histology
62
males
56
females
187
males
208
females
(
serum
chemistry
changes,
liver
histopathology)
Males
140
Females
16
Males
51
(
Liver
cell
vacuolization)

Females
2.3
(
Liver
cell
vacuolization)

Chu
et
al.

(
1982a)
Rat
SD
M
10
Drinking
water
0.7
8.5
80
28
days
Clinical
signs,

serum
biochemistry,

histology
80
­­
Not
modeled
­­

Condie
et
al.

(
1983)
Mouse
CD­
1
M
8­
16
Gavage
(
oil)
0
72
145
289
14
days
Serum
enzymes,

PAH
uptake
in
vitro,
histology
145
289
(
elevated
ALT,

decreased
PAH,

liver
and
kidney
histopathology)
73
34
(
Renal
mesangial
nephrosis)

Melnick
et
al.

(
1998)
Mouse
B6C3F
1
F
10
Gavage
(
oil)
0
200
500
3
weeks
(
5
d/
wk)
Body
and
liver
weights,
serum
chemistry,
liver
histology
200
500
(
liver
histopathology)
146*
104*

(
Liver
hydropic
degeneration)

Munson
et
al.

(
1982)
Mouse
CD­
1
M,
F
6­
12
Gavage
(
aqueous)
0
50
125
250
14
days
Body
and
organ
weights,
serum
chemistry,
hematology,

immune
function
125
250
(
elevated
serum
enzymes)
Not
modeled
­­

NTP
(
1989a)
Mouse
B6C3F
1
M
5
Gavage
(
oil)
0
50
100
200
400
600
14
days
Body
weight,

clinical
signs,

gross
pathology
200
400
(
stomach
nodules)
167
66
(
stomach
nodules)
Table
VIII­
15
(
cont.)

Reference
Species
Sex
n
Route
Dose
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)

Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
59
Coffin
et
al.

(
2000)
Mouse
B6C3F
1
F
10
Gavage
(
oil)
0
200
500
11
days
Relative
liver
weight;
liver
histopathology;

labeling
index
­­
200
(
liver
histopathology;

labeling
index)
Not
modeled
­­

NTP
(
1989a)
Rat
F344/
N
M,
F
5
Gavage
(
oil)
0
100
200
400
600
800
14
days
Body
weight,

clinical
signs,

gross
pathology
200
400
(
decreased
body
weight)
Not
modeled
­­

Ruddick
et
al.
(
1983)**
Rat
SD
F
14­

15
Gavage
(
oil)
0
50
100
200
Gestation
days
6­
15
Body
and
organ
weights;
maternal
serum
chemistry;

hematology,
and
histopathology;

developmental
parameters
50
100
(
sternebral
aberrations)
50
33
(
sternebral
aberrations)

*
Duration­
adjusted
dose
used
to
calculate
BMD
and
BMDL10
**
Ruddick
et
al
(
1983)
is
included
because
it
is
a
reproductive
study.

Abbreviations:
SD,
Sprague­
Dawley
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
60
Based
on
the
BMDL
10
identified
in
the
Aida
et
al.
(
1992a)
study,
the
Ten­
day
HA
for
a
10­
kg
child
is
calculated
according
to
the
following
equation:

Ten­
day
HA
=
(
2.3
mg/
kg­
day)(
10
kg)
=
0.23
mg/
L
(
rounded
to
0.2
mg/
L)
(
100)
(
1
L/
day)

where:

2.3
mg/
kg­
day
=
BMDL
10
based
on
the
occurrence
of
hepatic
vacuolization
in
female
rats
exposed
to
bromoform
in
the
diet
for
one
month
10
kg
=
Assumed
body
weight
of
a
child
100
=
Uncertainty
factor
based
on
NAS/
OW
guidelines.
Includes
a
factor
of
10
for
interspecies
variation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
When
the
BMDL
10
value
for
liver
cell
vacuolization
in
the
Aida
et
al.
(
1992a)
study
is
used,
the
Ten­
day
HA
for
a
10­
kg
child
is
calculated
to
be
0.2
mg/
L,
assuming
a
drinking
water
ingestion
rate
of
1
L/
day
and
use
of
a
composite
uncertainty
factor
of
100.
This
value
is
slightly
lower
than
the
Longer­
term
HA
for
a
10
kg
child
of
0.3
mg/
L
derived
using
subchronic
data
for
the
same
histopathological
endpoint.
This
small
difference
may
reflect
experimental
or
BMD
modeling
uncertainty.

For
purposes
of
comparison,
the
Ten­
day
HA
may
also
be
derived
using
the
conventional
NOAEL/
LOAEL
approach.
The
lowest
LOAEL
among
the
candidate
studies
was
100
mg/
kgday
for
developmental
effects
in
rats
(
Ruddick
et
al.,
1983).
The
NOAEL
in
this
study
was
50
mg/
kg­
day.
Aida
et
al.
(
1992a)
identified
NOAEL
values
of
56
and
62
mg/
kg­
day
and
LOAEL
values
of
187
and
208
mg/
kg­
day
for
histopathological
changes
in
male
and
female
rats
administered
bromoform
in
the
diet.
Chu
et
al.
(
1982a)
identified
a
freestanding
NOAEL
of
80
mg/
kg­
day
in
rats.
The
data
of
Aida
et
al.
(
1992a)
were
selected
for
calculation
of
the
Ten­
day
HA
because
the
study
tested
both
male
and
female
rats,
incorporated
more
dose
levels,
and
identified
both
NOAEL
and
LOAEL
values
and
because
the
NOAEL
identified
by
Chu
et
al.
(
1982a)
is
close
to
the
lowest
LOAEL
of
100
mg/
kg­
day.
Using
the
NOAEL
of
62
mg/
kg­
day
for
male
rats
and
assuming
the
default
body
weight
for
a
child
(
10
kg),
the
default
drinking
water
intake
for
a
child
(
1
L/
day),
and
a
composite
uncertainty
factor
of
100,
the
Ten­
day
HA
would
be
6.2
mg/
L
(
rounded
to
6
mg/
L).

c.
Longer­
term
Health
Advisory
Candidate
studies
for
derivation
of
the
Longer­
term
HA
are
summarized
in
Table
VIII­
16
below.
All
studies
identified
histopathological
changes
in
liver
tissue
as
the
critical
toxicological
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
61
Table
VIII­
16
Summary
of
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
for
Bromoform
Reference
Species
Sex
n
Route
Dose
Exposure
duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)*
BMDL
10
(
mg/
kg­
day)
*

Chu
et
al.

(
1982b)
Rat
SD
M,
F
20
Drinking
water
Male
0
0.65
6.1
57
218
Females
0
0.64
6.9
55
283
90
days
Body
weight,
serum
chemistry,
histology
57
218
(
decreased
weight
gain,
mild
hepatic
lesions)
Male
10
Females
No
fit
Male
5.9
(
Hepatic
lesions)

Females
­­

NTP
(
1989a)
Rat
F344/
N
M,
F
10
Gavage
(
corn
oil)
0
12
25
50
100
200
13
weeks
(
5
d/
wk)
Body
weight,

clinical
signs,
gross
necropsy,
histology
25
50
(
hepatic
vacuolization)
4.4
2.6
(
hepatic
vacuolization
in
male
rats)

NTP
(
1989a)
Mouse
B6C3F
1
M,
F
10
Gavage
(
corn
oil)
0
25
50
100
200
400
13
weeks
(
5
d/
wk)
Body
weight,

clinical
signs,
gross
necropsy,
histology
100
200
(
hepatic
vacuolization)
88
55
(
hepatic
vacuolization
in
male
mice)

Ruddick
et
al.
(
1983)**
Rat
SD
F
9­
14
Gavage
(
corn
oil)
0
50
100
200
Gestation
days
6­
15
Body
and
organ
weights;
maternal
serum
chemistry;

hematology,
and
histopathology;

developmental
parameters
50
100
(
sternebral
aberrations)
50
33
(
sternebral
aberrations)

NTP
(
1989b)
Mouse
ICR
Swiss
M,
F
20
Gavage
(
corn
oil)
0
50
100
200
105
days
Continuous
breeding
reprod.
study.
Body
and
organ
weights,

histopathology,

reproductive
parameters
100
200
(
decreased
maternal
body
weight)
­

(
not
modeled)
­

*
BMD
and
BMDL10
calculated
using
duration­
adjusted
doses
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
62
**
Ruddick
et
al
(
1983)
is
included
because
it
is
a
reproductive
study.

 
Not
modeled
Abbreviations:
SD,
Sprague­
Dawley
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
63
effect.
In
one
NTP
(
1989a)
study,
F344/
N
rats
were
administered
bromoform
by
gavage
at
doses
ranging
from
12
to
200
mg/
kg­
day
for
5
days/
week
for
13
weeks.
This
study
identified
a
NOAEL
of
25
mg/
kg­
day
and
a
LOAEL
of
50
mg/
kg­
day
based
on
hepatic
vacuolization
observed
in
male
rats.
BMD
modeling
with
the
BMDS
program
calculated
a
duration­
adjusted
BMD
of
4.4
mg/
kg­
day
(
based
on
duration­
adjusted
doses),
with
a
corresponding
BMDL
10
of
2.6
mg/
kg­
day.
These
values
were
the
lowest
among
the
candidate
studies.

In
an
analogous
subchronic
oral
exposure
study,
NTP
(
1989a)
exposed
mice
of
both
sexes
to
doses
of
bromoform
ranging
from
25
to
400
mg/
kg­
day
in
addition
to
the
control.
This
study
identified
NOAEL
and
LOAEL
values
of
100
and
200
mg/
kg­
day,
respectively,
based
on
hepatic
vacuolization.
BMD
modeling
with
the
BMDS
program
calculated
a
BMD
of
88
mg/
kg­
day
(
based
on
duration­
adjusted
doses),
with
a
corresponding
BMDL
10
of
55
mg/
kg­
day.
These
values
were
approximately
20­
fold
higher
than
the
BMD
and
BMDL
10
calculated
for
the
NTP
(
1989a)
oral
exposure
study
in
rats.

Chu
et
al.
(
1982b)
exposed
rats
of
both
sexes
to
bromoform
in
the
drinking
water
for
90
days.
The
doses
of
bromoform
ranged
from
0.64
to
283
mg/
kg­
day
in
addition
to
the
control.
This
study
identified
NOAEL
and
LOAEL
values
of
57
and
218
mg/
kg­
day
,
respectively,
based
on
decreased
weight
gain
and
mild
hepatic
lesions
in
male
mice.
BMD
modeling
identified
BMD
and
BMDL
10
values
of
10
and
5.9
mg/
kg­
day
using
data
for
occurrence
of
hepatic
lesions
in
male
mice.
These
values
were
approximately
two­
fold
higher
than
the
BMD
and
BMDL
10
values
derived
using
data
from
the
NTP
(
1989a)
oral
exposure
study
in
rats.
Strengths
of
this
study
include
exposure
via
drinking
water,
larger
sample
size
(
20
animals/
treatment
group),
and
the
administration
of
lower
doses
than
used
in
the
NTP
(
1989a)
subchronic
studies.
Liver
histopathology
data
from
this
study
were
reported
as
combined
lesions,
with
the
types
of
lesions
described
in
the
text.

The
NTP
(
1989a)
oral
exposure
study
conducted
in
rats
was
selected
for
the
derivation
of
the
Longer­
term
HA
on
the
basis
of
the
low
values
obtained
for
the
BMD
and
BMDL
10.
Selection
of
this
study
is
strongly
supported
by
the
results
of
Chu
et
al.
(
1982b),
which
identified
slightly
higher
values
in
a
drinking
water
study.

Based
on
the
BMDL
10
identified
in
the
NTP
(
1989a)
rat
study,
the
Longer­
term
HA
for
a
10­
kg
child
is
calculated
according
to
the
following
equation:

Longer­
term
HA
=
(
2.6
mg/
kg­
day)(
10
kg)
=
0.26
mg/
L
(
rounded
to
0.3
mg/
L)
(
100)
(
1
L/
day)

where:

2.6
mg/
kg­
day
=
Duration­
adjusted
BMDL
10
based
on
the
occurrence
of
hepatic
vacuolization
in
male
rats
exposed
to
bromoform
by
gavage
for
13
weeks
10
kg
=
Assumed
body
weight
of
a
child
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64
100
=
Uncertainty
factor
based
on
NAS/
OW
guidelines;
includes
a
factor
of
10
for
interspecies
variation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
1
L/
day
=
Assumed
water
consumption
of
a
10­
kg
child
The
Longer­
term
HA
for
an
adult
consuming
2
liters
of
water
per
day
is
calculated
according
to
the
following
equation:

Longer­
term
HA
=
(
2.6
mg/
kg­
day)(
70kg)
=
0.91
mg/
L
(
rounded
to
0.9
mg/
L)
(
100)
(
2
L/
day)

where:

2.6
mg/
kg­
day
=
Duration­
adjusted
BMDL
10
based
on
the
occurrence
of
hepatic
vacuolization
in
male
rats
exposed
to
bromoform
by
gavage
for
13
weeks
70
kg
=
Assumed
body
weight
of
an
adult
100
=
Composite
uncertainty
factor
based
on
NAS/
OW
guidelines;
includes
a
factor
of
10
for
interspecies
variation
and
a
factor
of
10
for
protection
of
sensitive
human
populations
2
L/
day
=
Assumed
water
consumption
of
a
70­
kg
adult
For
purposes
of
comparison,
the
Longer­
term
Health
Advisories
may
also
be
derived
using
the
conventional
NOAEL/
LOAEL
approach.
The
chronic
oral
exposure
study
conducted
by
NTP
(
1989a)
identified
a
NOAEL
of
25
mg/
kg­
day
based
on
the
absence
of
clinical
signs
or
histological
alterations
in
rats
exposed
to
bromoform
for
13
weeks.
Using
this
value
and
assuming
default
body
weights
(
10
and
70
mg/
kg­
day,
for
adults
and
children,
respectively),
default
drinking
water
intake
rates
(
1
and
10
L/
day
for
children
and
adults,
respectively),
a
composite
uncertainty
factor
of
100,
and
an
exposure
duration
factor
of
5/
7,
the
Longer­
term
HAs
for
a
child
and
an
adult
would
be
2
mg/
L
and
6
mg/
L,
respectively.

d.
Reference
Dose,
Drinking
Water
Equivalent
Level
and
Lifetime
Health
Advisory
This
section
reports
the
existing
RfD
value
for
bromoform
and
describes
the
derivation
of
the
RfD
for
this
compound.
This
section
also
describes
the
calculation
of
Drinking
Water
Equivalent
Level
and
Lifetime
Health
Advisory
values
which
require
the
RfD
as
input.
For
this
document,
new
and
existing
studies
were
reviewed
and
appropriate
candidate
data
were
selected
for
benchmark
(
BMD)
dose
modeling.
The
results
of
BMD
modeling
were
used
in
conjunction
with
appropriate
uncertainty
factors
to
calculate
the
RfD.
A
comparison
of
the
RfD
derived
using
Draft
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VIII
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65
the
BMD
approach
to
the
results
obtained
using
the
conventional
NOAEL/
LOAEL
approach
is
also
provided.

Description
of
the
Existing
RfD
The
existing
RfD
for
bromoform
is
0.02
mg/
kg­
day
(
IRIS,
1993b).
This
value
was
derived
using
a
duration­
adjusted
NOAEL
of
17.9
mg/
kg­
day
identified
for
the
occurrence
of
hepatic
lesions
in
F344
rats
administered
dibromochloromethane
by
corn
oil
gavage
5
days/
week
for
13
weeks
(
NTP,
1989a).
An
uncertainty
factor
of
1000
was
used
to
account
for
extrapolation
from
animal
data,
for
protection
of
sensitive
human
subpopulations,
and
for
use
of
a
subchronic
study.
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VIII
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66
Identification
of
Candidate
Studies
for
the
Derivation
of
the
RfD
Three
chronic
exposure
studies,
a
subchronic
exposure
study,
a
prenatal
developmental
toxicity
study,
and
a
reproductive
toxicity
study
were
considered
for
derivation
of
the
RfD
for
bromoform.
These
studies
are
summarized
in
Table
VIII­
17
(
below).

Tobe
et
al.
(
1982)
administered
bromoform
microencapsulated
in
the
diet
to
Wistar
rats
at
dose
levels
ranging
from
22
to
619
mg/
kg­
day
for
24
months.
A
NOAEL
of
22
mg/
kg­
day
and
a
LOAEL
of
90
mg/
kg­
day
were
identified
based
on
gross
liver
lesions
and
changes
in
clinical
chemistry
parameters
in
male
rats.

NTP
(
1989a)
conducted
chronic
oral
exposure
studies
in
rats
and
mice.
In
the
rat
study,
animals
were
administered
bromoform
by
gavage
in
oil
for
5
days/
week
for
103
weeks
at
doses
of
100
or
200
mg/
kg­
day.
This
study
identified
the
low
dose
as
the
LOAEL
based
on
histologic
lesions
in
the
liver
(
fatty
change
and
chronic
inflammation).
In
the
mouse
study,
animals
were
administered
bromoform
by
gavage
in
corn
oil,
5
days/
week
for
103
weeks
at
doses
of
50
or
100
mg/
kg­
day
for
male
mice
and
100
or
200
mg/
kg­
day
for
female
mice.
Although
no
treatmentrelated
effects
were
observed
in
the
male
mice
at
the
dose
levels
tested,
treatment­
related
histologic
lesions
in
the
liver
(
fatty
changes)
were
observed
in
both
low­
and
high­
dose
females.
Accordingly,
this
study
identified
a
LOAEL
of
100
mg/
kg­
day
in
female
mice.

A
subchronic
oral
exposure
study
conducted
in
rats
(
NTP,
1989a)
was
also
considered
as
a
candidate
for
derivation
of
the
RfD.
This
study
utilized
five
doses
of
bromoform
ranging
from
12
to
200
mg/
kg­
day
in
addition
to
the
control.
The
compound
was
administered
to
ten
animals
per
treatment
group
by
gavage
in
corn
oil,
5
days
per
week
for
13
weeks.
The
endpoints
evaluated
included
body
weight,
clinical
signs,
gross
necropsy,
and
histological
changes.
This
study
identified
a
NOAEL
and
LOAEL
of
25
and
50
mg/
kg­
day,
respectively,
on
the
basis
of
histopathological
changes
(
vacuolization)
in
the
liver.
The
LOAEL
identified
in
this
study
was
the
lowest
among
all
candidate
studies.

The
developmental
study
conducted
by
Ruddick
et
al.
(
1983)
identified
NOAEL
and
LOAEL
values
of
50
and
100
mg/
kg­
day,
respectively,
for
sternebral
variations
in
the
offspring
of
female
rats
dosed
with
bromoform
on
gestation
days
6­
15.
The
reproductive
toxicity
study
reported
by
NTP
(
1989b)
identified
NOAEL
and
LOAEL
values
of
100
and
200
mg/
kg­
day
for
reduced
maternal
body
weight
and
decreased
postnatal
survival
and
liver
histopathology
in
F
1
mice
of
both
sexes.

Method
of
Analysis
Selected
data
from
the
candidate
studies
were
analyzed
using
the
benchmark
dose
(
BMD)
modeling
approach.
Initially,
data
sets
for
potentially
sensitive
endpoints
were
selected
as
described
in
U.
S.
EPA
(
1998b)
and
analyzed
using
the
Crump
Benchmark
Dose
Modeling
Software
(
K.
S.
Crump,
Inc.).
Following
the
release
of
Version
1.2
of
the
BMDS
program
(
U.
S.
EPA,
2000a),
data
from
the
NTP
(
1989a)
subchronic
and
chronic
studies
conducted
in
rats
were
reanalyzed
in
accordance
with
proposed
U.
S.
EPA
(
2000b)
recommendations.
An
advantage
of
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VIII
­
67
analysis
with
the
BMDS
software
is
that
several
additional
models
are
available
to
fit
the
data.
The
results
of
the
analysis
using
the
BMDS
software
are
included
in
Table
VIII­
15.
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68
Table
VIII­
17
Summary
of
Candidate
Studies
for
Derivation
of
the
RfD
for
Bromoform
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
BMD
(
mg/
kg­
day)
BMDL
10
(
mg/
kg­
day)
*

Tobe
et
al.

(
1982)
Rat
Wistar
M,
F
40
Male
0
22
90
364
Female
0
38
152
619
Diet
24
months
Body
weight,
serum
chemistry,
gross
pathology
22
90
(
serum
chemistry
changes,
gross
liver
lesions)
Not
modeled
­­

NTP
(
1989a)
Rat
F344/
N
M,
F
10
0
12
25
50
100
200
Gavage
(
corn
oil)
13
weeks
(
5
days/
wk)
Body
weight,
clinical
signs,
gross
necropsy,

histology
25
50
(
hepatic
vacuolization)
4.4
2.6
(
Hepatic
vacuolization
in
male
rats)

NTP
(
1989a)
Rat
F344/
N
M,
F
50
0
100
200
Gavage
(
corn
oil)
103
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
gross
necropsy,

histology
­­
100
(
decreased
body
weight,
lethargy,

mild
liver
histopathology)
13
1.4
(
fatty
changes
in
liver
of
males)

NTP
(
1989a)
Mouse
B6C3F
1
M,
F
50
Male
0
50
100
Female
0
100
200
Gavage
(
corn
oil)
103
weeks
(
5
d/
wk)
Body
weight,
clinical
signs,
gross
necropsy,

histology
100
(
male)
100
(
female)

(
decreased
body
weight,
mild
liver
histopathology)
14.2*
10.6*

(
fatty
changes
in
liver
of
females)

Ruddick
et
al.

(
1983)*
Rat
SD
F
9­
14
0
50
100
200
Gavage
(
con
oil)
Gestation
days
6­
15
Developmental
toxicity
study;
body
and
organ
weights
50
(
developmental)

200
(
maternal)
100
(
sternebral
variations)
50
33
(
sternebral
variations)
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69
NTP
(
1989b)*
Mouse
B6C3F
1
M,
F
50
0
50
100
200
Gavage
(
corn
oil)
105
days
Continuous
breeding
reprod.
study.
Body
and
organ
weights,
histopathology,

reproductive
parameters
100
200
(
decreased
maternal
body
weight;

reduced
postnatal
survival
and
liver
histopathology
in
F
1
generation
males
and
females)
­

(
not
modeled)
­

*
BMD
and
BMDL10
values
were
derived
using
duration­
adjusted
doses.

*
BMD
modeled
using
the
Crump
BMD
software
(
K.
S.
Crump,
Inc.)

*
These
studies
are
included
because
they
are
reproductive/
developmental
studies.

Abbreviations:
SD,
Sprague­
Dawley
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70
Choice
of
Principal
Study
and
Critical
Effect
for
the
RfD
The
subchronic
study
conducted
by
NTP
(
1989a)
was
selected
for
derivation
of
the
RfD.
Two
factors
supported
selection
of
this
study.
First,
the
critical
effect
identified
in
the
subchronic
study
was
consistent
with
the
critical
effects
identified
in
the
chronic
NTP
studies
(
fatty
changes
in
liver
of
male
rats
and
female
mice).
Second,
BMD
modeling
of
both
data
sets
supported
selection
of
the
subchronic
study.
BMD
analysis
using
the
BMDS
program
calculated
a
BMD
of
13
mg/
kg­
day
for
fatty
changes
in
the
liver
of
males
in
the
chronic
study,
with
a
corresponding
BMDL
10
of
1.4
mg/
kg­
day.
The
lowest
duration­
adjusted
dose
in
this
study
was
71
mg/
kg­
day
and
the
response
at
this
dose
was
high
(
49/
50).
The
magnitude
of
the
difference
between
the
BMD
and
BMDL
10
values
thus
reflects
considerable
uncertainty
about
the
shape
of
the
curve
in
the
low­
dose
region.

Duration­
adjusted
BMD
and
BMDL
10
values
for
hepatic
vacuolization
in
male
rats
of
4.4
and
2.6
mg/
kg­
day,
respectively,
were
obtained
using
data
for
the
subchronic
study.
This
BMD
is
approximately
three­
fold
less
than
the
BMD
calculated
from
chronic
data
(
above).
The
availability
of
response
data
for
three
doses
below
71
mg/
kg­
day
(
duration­
adjusted
dose)
in
the
subchronic
study
provided
additional
information
about
the
shape
of
the
dose­
response
curve
in
the
region
of
interest,
and
thus
a
more
reliable
estimate
of
the
BMD.
Although
the
BMDL
10
value
for
the
subchronic
study
is
higher
than
the
value
for
the
chronic
study,
this
observation
reflects
less
uncertainty
(
smaller
confidence
interval)
in
the
estimate
of
the
subchronic
BMD
when
the
results
for
the
two
studies
are
compared.
The
BMDL
10
value
calculated
from
the
subchronic
NTP
(
1987)
data
was
therefore
selected
for
derivation
of
the
RfD
for
bromoform.

The
remaining
studies
were
eliminated
from
consideration
for
the
following
reasons.
The
study
conducted
by
Tobe
et
al.
(
1982)
did
not
identify
a
suitably
sensitive
endpoint
(
histopathological
examination
was
not
conducted)
and
the
data
were
never
formally
published
or
submitted
for
peer
review.
The
chronic
study
conducted
by
NTP
(
1989a)
in
mice
reported
mild
histopathological
changes
in
female
mice
at
a
duration­
adjusted
dose
of
71
mg/
kg­
day,
the
lowest
tested
in
this
gender.
However,
both
the
BMD
and
BMDL
10
were
higher
than
those
identified
in
the
subchronic
study
in
rats,
and
these
values
were
considered
less
reliable
in
the
absence
of
data
at
lower
doses.
The
LOAELs
identified
in
the
developmental
(
Ruddick
et
al.,
1983)
and
reproductive
(
NTP,
1989a)
studies
were
higher
than
the
LOAEL
values
observed
for
histopathology
effects
in
the
NTP
subchronic
study.
The
BMD
and
BMDL
10
calculated
for
sternebral
variations
in
the
Ruddick
et
al.
(
1983)
were
approximately
10­
fold
higher
than
those
identified
in
the
subchronic
study.

Derivation
of
the
RfD
The
duration­
adjusted
BMDL
10
from
the
subchronic
NTP
(
1989a)
rat
study
was
selected
for
derivation
of
the
RfD
for
bromoform.
The
RfD
is
calculated
using
the
following
equation:

RfD
=
(
2.6
mg/
kg­
day)
=
0.03
mg/
kg­
day
100
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VIII
­
71
where:

2.6
mg/
kg­
day
=
Duration­
adjusted
BMDL
10
based
on
hepatocellular
vacuolization
in
the
liver
of
male
rats
100
=
Composite
uncertainty
factor
based
on
NAS/
OW
guidelines;
includes
a
factor
of
10
for
interspecies
variation,
a
factor
of
10
for
protection
of
sensitive
human
populations
A
composite
uncertainty
factor
of
100
was
used
in
the
calculation
of
the
bromoform
RfD.
The
standard
factors
of
10
were
used
for
interspecies
extrapolation
and
for
protection
of
sensitive
subpopulations.
No
uncertainty
factor
was
added
for
extrapolation
from
a
subchronic
to
a
chronic
study
because
the
BMD
and
BMDL
10
for
the
subchronic
study
was
either
comparable
to
or
lower
than
the
corresponding
values
from
the
chronic
study.
This
observation
suggests
that
a
cumulative
effect
on
the
liver
does
not
occur
for
the
endpoints
examined.
The
database
for
bromoform
includes
subchronic
and
chronic
bioassays
conducted
in
rats
and
mice
(
e.
g.
NTP
1989a),
a
two­
generation
reproductive
toxicity
study
in
mice
(
NTP
1989b),
and
a
developmental
toxicity
study
in
rats
(
Ruddick
et
al.
1983).
Therefore,
the
database
for
bromoform
was
considered
sufficient
and
an
uncertainty
factor
for
database
deficiencies
was
not
included
in
the
calculation.

The
DWEL
for
bromoform
is
calculated
as
follows:

DWEL
=
(
0.03
mg/
kg­
day)
(
70
kg)
=
1.0
mg/
L
(
1000
µ
g/
L)
2
L/
day
where:

0.03
mg/
kg­
day
=
RfD
70
kg
=
assumed
weight
of
an
adult
2
L/
day
=
assumed
water
consumption
by
a
70­
kg
adult
Lifetime
Health
Advisory
The
Lifetime
Health
Advisory
(
HA)
represents
that
portion
of
an
individual's
total
exposure
that
is
attributed
to
drinking
water
and
is
considered
protective
of
noncarcinogenic
health
effects
over
a
lifetime
of
exposure.
Bromoform
has
been
categorized
with
respect
to
carcinogenic
potential
as
Group
B2:
Probable
human
carcinogen
(
IRIS,
1993b).
Therefore,
in
accordance
with
U.
S.
EPA
Policy,
a
Lifetime
Health
Advisory
is
not
recommended.
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2003
VIII
­
72
Alternative
Approach
for
Derivation
of
the
RfD
For
comparison,
the
RfD
can
be
calculated
using
the
conventional
NOAEL/
LOAEL
approach.
The
subchronic
NTP
(
1989a)
oral
exposure
study
identified
a
NOAEL
of
25
mg/
kgday
based
on
absence
of
histopathological
effects
in
rats
exposed
to
dietary
bromoform
for
13
weeks.
Using
this
value,
a
duration
adjustment
factor
of
5/
7,
and
an
uncertainty
factor
of
1000
(
including
factors
of
10
for
interspecies
extrapolation,
protection
of
sensitive
subpopulations,
and
use
of
a
subchronic
study),
the
RfD
would
be
0.02
mg/
kg­
day.
The
corresponding
DWEL
would
be
0.7
mg/
L
assuming
an
adult
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L/
day.

2.
Carcinogenic
Effects
a.
Categorization
of
Carcinogenic
Potential
Previous
Evaluations
of
Carcinogenic
Potential
The
Carcinogenic
Risk
Assessment
Verification
Endeavor
(
CRAVE)
group
of
the
U.
S.
EPA
has
reviewed
the
available
evidence
on
the
carcinogenicity
of
bromoform
and
has
assigned
it
to
Group
B2:
probable
human
carcinogen
(
IRIS,
1993b).
Assignment
to
this
category
is
appropriate
for
chemicals
where
there
are
no
or
inadequate
human
data,
but
which
have
sufficient
animal
data
to
indicate
carcinogenic
potential.

IARC
(
1999b)
has
recently
re­
evaluated
the
carcinogenic
potential
of
bromoform.
IARC
concluded
that
there
is
limited
evidence
of
carcinogenicity
in
experimental
animals
and
inadequate
evidence
in
humans
for
bromoform.
Bromoform
is
therefore
categorized
as
Group
3:
not
classifiable
as
to
carcinogenicity
in
humans.

Categorization
of
Carcinogenic
Potential
Under
the
Proposed
1999
Cancer
Guidelines
Cancer
Hazard
Summary
Under
the
proposed
guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
bromoform
is
likely
to
be
carcinogenic
to
humans
by
all
routes
of
exposure.
This
descriptor
is
appropriate
when
the
available
tumor
data
and
other
key
data
are
adequate
to
demonstrate
carcinogenic
potential
to
humans.
This
finding
is
based
on
the
weight
of
experimental
evidence
in
animal
models
which
shows
carcinogenicity
by
modes
of
action
that
are
relevant
to
humans.

Supporting
Information
for
Cancer
Hazard
Assessment
Human
Data
The
information
on
the
carcinogenicity
of
bromoform
from
human
studies
is
inadequate.
There
are
no
epidemiological
data
specifically
relating
increased
incidence
of
cancer
to
exposure
to
bromoform.
There
are
equivocal
epidemiological
data
describing
a
weak
association
of
chlorinated
drinking
water
exposures
with
increased
incidences
of
bladder,
rectal,
and
colon
Draft
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VIII
­
73
cancer.
U.
S.
EPA
has
determined
that
these
studies
cannot
attribute
the
observed
effects
to
a
single
compound,
as
chlorinated
water
contains
numerous
other
disinfection
byproducts
that
are
potentially
carcinogenic.

Animal
Data
The
carcinogenicity
of
bromoform
has
been
investigated
in
two
species.
These
studies
include
a
well­
designed
and
conducted
corn
oil
gavage
study
conducted
in
rats
and
mice
and
a
study
in
which
male
Strain
A
mice
were
administered
bromoform
by
intraperitoneal
injection.
No
data
are
available
on
the
carcinogenic
potential
of
bromoform
administered
via
the
inhalation
or
dermal
routes.

In
the
corn
oil
gavage
study
(
NTP,
1989a),
neoplasms
of
the
large
intestine
(
adenomatous
polyps
or
adenocarcinoma)
were
observed
in
male
and
female
rats.
The
response
for
combined
adenoma
and
carcinoma
reached
statistical
significance
in
female
rats.
The
occurrence
of
tumors
of
the
large
intestine
in
this
study
was
considered
biologically
significant
because
they
are
historically
rare
in
rats.
NTP
(
1989a)
concluded
that
there
was
clear
evidence
for
carcinogenicity
in
females
and
some
evidence
of
carcinogenicity
in
males.
No
evidence
of
bromoform
carcinogenicity
was
observed
in
male
or
female
mice.
Intraperitoneal
injection
of
Strain
A
mice
with
three
concentrations
of
bromoform
resulted
in
significantly
increased
tumor
incidence
only
at
the
middle
dose
tested.

Structural
Analogue
Data
Trihalomethanes
structurally
related
to
bromoform
have
shown
varying
degrees
of
carcinogenic
potential
in
rodents.
Chloroform,
the
most
extensively
characterized
trihalomethane,
is
reported
to
be
carcinogenic
at
high
doses
in
several
chronic
animal
bioassays,
with
significant
increases
in
the
incidence
of
liver
tumors
in
male
and
female
mice
and
significant
increases
in
the
incidence
of
kidney
tumors
in
male
rats
and
mice
(
U.
S.
EPA,
2001).
The
occurrence
of
tumors
in
animals
exposed
to
chloroform
is
demonstrably
species­,
strain­,
and
gender­
specific,
and
has
only
been
observed
under
dose
conditions
that
caused
cytotoxicity
and
regenerative
cell
proliferation
in
the
target
organ.
The
cancer
database
for
structurally­
related
brominated
trihalomethanes
is
more
limited,
but
includes
well­
conducted
studies
performed
by
the
National
Toxicology
Program.
In
a
two­
year
corn
oil
gavage
study
of
bromodichloromethane,
NTP
(
1987)
found
clear
evidence
for
carcinogenicity
in
male
and
female
rats
(
large
intestine
and
kidney)
and
male
(
kidney)
and
female
(
liver)
mice.
Tumasonis
et
al.
(
1987)
reported
a
statistically
significant
increase
in
the
incidence
of
hepatic
neoplastic
nodules
in
rats
exposed
to
bromodichloromethane
in
the
drinking
water.
In
a
two­
year
corn
oil
gavage
study
of
dibromochloromethane,
NTP
(
1985)
determined
that
there
was
some
evidence
of
carcinogenicity
in
female
mice
and
equivocal
evidence
of
carcinogenicity
in
male
mice,
based
on
the
occurrence
of
hepatocellular
adenomas
and
carcinomas.
Other
oral
exposure
studies
found
no
evidence
for
carcinogenicity
of
bromodichloromethane
(
Aida
et
al.,
1992b)
or
dibromochloromethane
(
Tobe
et
al.,
1982;
Voronin
et
al.,
1987).

Other
Key
Data
Draft
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2003
VIII
­
74
Bromoform
is
formed
as
a
byproduct
of
drinking
water
disinfection
with
chlorine.
Exposure
to
bromoform
may
occur
via
ingestion
of
tap
water,
via
dermal
contact
during
showering
or
bathing,
or
by
inhalation
of
bromoform
volatilized
during
household
activities.
Absorption
of
single
oral
doses
appears
to
be
extensive.
Bromoform
is
rapidly
metabolized
and
eliminated
predominately
as
expired
volatiles,
CO
2,
or
CO.
Only
a
small
amount
(
less
than
10%)
is
eliminated
in
urine
or
in
feces.
No
comprehensive
tissue
data
are
available
regarding
the
bioaccumulation
or
retention
of
bromoform
following
repeated
exposure.
However,
because
of
the
rapid
metabolism
and
excretion
of
bromoform,
marked
accumulation
and
retention
is
not
expected.

Bromoform
itself
is
not
directly
reactive
with
DNA.
Metabolism
to
reactive
species
is
a
prerequisite
for
toxicity,
as
inferred
from
metabolic
induction
and
inhibition
studies.
In
vitro
and
in
vivo
studies
of
the
mutagenic
and
genotoxic
potential
of
bromoform
have
yielded
both
positive
and
negative
results.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid)
and,
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
However,
because
a
majority
of
studies
yielded
positive
results,
bromoform
is
considered
to
be
at
least
weakly
mutagenic
and
genotoxic.
Recent
studies
conducted
with
strains
of
Salmonella
engineered
to
express
rat
theta­
class
glutathione
Stransferase
suggest
that
mutagenicity
of
the
brominated
trihalomethanes
may
be
mediated
by
glutathione
conjugation.

Mode
of
Action
The
mode
of
action
for
tumor
induction
by
bromoform
has
not
been
clearly
elucidated
and
may
involve
contributions
from
multiple
bioactivation
pathways.
In
each
case,
toxicity
is
believed
to
result
from
interaction
of
reactive
metabolites
with
cellular
macromolecules.
Proposed
bioactivation
pathways
for
bromoform
include:
1)
production
of
reactive
dihalocarbonyls
by
oxidative
metabolism;
2)
production
of
reactive
dihalomethyl
radicals
by
oxidative
metabolism;
and
3)
formation
of
DNA­
reactive
species
via
a
glutathione­
dependent
pathway.
The
relative
contribution
of
each
pathway
to
tumor
induction
by
bromoform
has
not
been
characterized.
It
is
possible
that
only
the
latter
two
processes
lead
to
DNA
damage
in
vivo,
because
the
highly
reactive
dihalocarbonyl
intermediate
may
not
survive
long
enough
to
enter
the
nucleus
and
react
with
DNA.
For
this
reason,
cytotoxicity
may
be
the
primary
consequence
of
the
oxidative
pathway.
Cytotoxicity
coupled
with
regenerative
hyperplasia
is
considered
the
primary
mode
of
action
for
tumor
formation
following
exposure
to
high
concentrations
of
chloroform,
a
structurally­
related
trihalomethane
which
has
low
genotoxic
potential.
Data
to
support
a
similar
primary
mode
of
action
for
tumor
development
in
liver,
kidney,
and
large
intestine
are
currently
lacking
for
bromoform.
In
the
absence
of
such
information,
combined
with
a
positive
weight­
ofevidence
evaluation
for
genotoxicity,
the
mode
of
action
for
tumor
development
is
assumed
to
be
a
linear
process.
The
processes
leading
to
tumor
formation
in
animals
are
expected
to
be
relevant
to
humans.

Conclusion
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2003
VIII
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75
Under
the
proposed
guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999)
bromoform
is
likely
to
be
carcinogenic
to
humans
by
the
oral
route.
This
weight­
of­
evidence
evaluation
is
based
on
1)
observations
of
tumors
in
rats
treated
by
oral
pathways;
2)
lack
of
epidemiological
data
specific
to
bromoform
and
equivocal
data
for
drinking
water
drinking
water
exposures
that
cannot
reliably
be
attributed
to
bromoform
among
multiple
other
disinfection
byproducts;
3)
positive
results
for
a
majority
of
the
available
genotoxicity
and
mutagenicity
tests;
and
4)
metabolism
and
mode
of
action
that
are
reasonably
expected
to
be
comparable
across
species.
Although
no
cancer
data
exist
for
exposures
via
the
dermal
or
inhalation
pathways,
the
weight­
of­
evidence
conclusion
is
considered
to
be
applicable
to
these
pathways
as
well.
The
finding
for
inhalation
is
based
on
the
observation
that
patterns
of
metabolizing
enzyme
activity
in
male
rats
are
similar
following
exposure
to
a
structurally­
related
compound
(
bromodichloromethane)
via
the
inhalation
and
gavage
routes.
Bromoform
absorbed
through
the
skin
is
expected
to
be
metabolized
and
cause
toxicity
in
much
the
same
way
as
bromoform
absorbed
by
the
oral
and
inhalation
routes.

b.
Choice
of
Study
for
Quantification
of
Carcinogenic
Risk
A
single
oral
exposure
study
was
available
for
the
quantification
of
carcinogenic
risk
associated
with
oral
exposure
to
bromoform.
NTP
(
1989a)
conducted
an
oral
exposure
study
in
B6C3F
1
mice
and
F344/
N
rats.
No
evidence
of
carcinogenicity
was
observed
in
male
B6C3F
1
mice
exposed
to
bromoform
via
gavage
(
corn
oil)
at
doses
up
to
100
mg/
kg­
day,
or
in
female
mice
exposed
at
doses
up
to
200
mg/
kg­
day
for
5
days/
week.
Male
and
female
F344/
N
rats
(
50
rats/
sex/
dose)
were
administered
bromoform
via
gavage
at
doses
of
0,
100,
or
200
mg/
kg­
day
for
5
days/
week
for
103
weeks
(
NTP
1989a).
At
termination,
all
animals
were
necropsied,
and
a
thorough
histological
examination
of
tissues
was
performed.
Adenomatous
polyps
or
adenocarcinomas
of
the
large
intestine
were
noted
in
three
high­
dose
male
rats,
eight
high­
dose
female
rats,
and
one
low­
dose
female
rat
(
Table
VIII­
18).
Despite
the
small
number
of
tumors
found,
the
increase
was
considered
biologically
significant
because
these
tumors
are
historically
rare
in
the
rat.
The
study
authors
concluded
that
there
was
some
evidence
for
carcinogenic
activity
in
male
rats
and
clear
evidence
in
female
rats.

c.
Extrapolation
model
The
LMS
model
(
U.
S.
EPA,
1986)
and
the
default
linear
approach
described
by
the
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1996;
1999)
were
used
to
quantify
the
risk
associated
with
exposure
to
bromoform.
Although
data
are
mixed
,
U.
S.
EPA
has
previously
concluded
that
the
weight
of
evidence
suggests
that
bromoform
is
mutagenic
(
see
Section
V.
F.
3).
At
the
present
time,
there
are
no
data
which
indicate
that
bromoform­
induced
tumorigenesis
occurs
as
a
consequence
of
cytotoxicity
followed
by
regenerative
hyperplasia.
Thus,
use
of
a
linear
approach
was
considered
appropriate
for
quantification
of
cancer
risk
associated
with
exposure
to
bromodichloromethane.

Table
VIII­
18
Tumor
Frequencies
in
Rats
Exposed
to
Bromoform
in
Corn
Oil
for
2
Years
­
Adapted
from
NTP
(
1989a)
Draft
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VIII
­
76
Animal
Tissue/
Tumor
Tumor
Frequency
Control
100
mg/
kg
200
mg/
kg
Male
rat
Large
intestine
Adenocarcinoma
0/
50
0/
50
1/
50
Polyp
(
adenomatous)
0/
50
0/
50
2/
50
Female
rat
Large
intestine
Adenocarcinoma
0/
48
0/
50
2/
50
Polyp
(
adenomatous)
0/
48
1/
50
6/
50
d.
Cancer
Potency
and
Unit
Risk
Estimates
of
cancer
risk
associated
with
exposure
to
bromoform
are
summarized
in
Table
VIII­
19.
U.
S.
EPA
(
1994b)
reported
a
cancer
potency
estimate
of
7.9
x
10­
3
(
mg/
kg­
day)­
1
for
bromoform
based
on
the
incidence
of
intestinal
tumors
in
rats
and
derived
using
recommendations
in
the
1986
Cancer
Guidelines
(
U.
S.
EPA,
1986).
The
calculated
value
for
unit
risk
is
2.3
×
10­
7
(
µ
g/
L)­
1.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
40
µ
g/
L
associated
with
a
10­
5
risk.

A
cancer
potency
estimate
of
4.6x10­
3
(
mg/
kg­
day)­
1
(
U.
S.
EPA,
1998b)
based
on
the
incidence
of
intestinal
tumors
in
rats
was
calculated
using
the
LMS
model
and
a
scaling
factor
of
body
weight3/
4.
Use
of
this
scaling
factor
is
consistent
with
recommendations
in
U.
S.
EPA
(
1992b).
Unit
risk
was
estimated
for
bromoform
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
The
calculated
value
for
unit
risk
is
1.30
×
10­
7
(
µ
g/
L)­
1.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
77
µ
g/
L
associated
with
a
10­
5
risk
(
8
µ
g/
L
for
a
risk
of
10­
6).

Cancer
risk
estimates
were
also
obtained
using
the
LED
10
(
the
lower
95%
confidence
limit
on
a
dose
associated
with
10%
extra
risk)
for
hepatic
tumors
and
assuming
a
linear
mode
of
action
for
the
carcinogenicity
of
bromoform
(
Table
VIII­
19).
A
cancer
potency
value
of
4.5
x
10­
3
(
mg/
kg­
day)­
1
was
derived
using
this
approach.
A
unit
risk
of
1.3
x
10­
7
(
µ
g/
L)­
1
was
calculated
using
an
assumed
body
weight
of
70
kg
and
a
drinking
water
ingestion
rate
of
2
L.
This
estimate
was
used
to
calculate
a
drinking
water
concentration
of
78
µ
g/
L
associated
with
a
10­
5
risk
(
8
µ
g/
L
for
10­
6
risk).
These
values
are
similar
to
values
derived
using
the
LMS
approach
with
body
weight
scaling
to
the
3/
4
power.

There
are
no
data
to
suggest
that
tumor
incidence
in
the
large
intestine
is
influenced
by
the
use
of
an
oil
vehicle.
Therefore,
the
risk
estimates
reported
above
are
believed
to
be
applicable
to
drinking
water
exposures.

Table
VIII­
19
Carcinogenic
Risk
Estimates
for
Bromoform
Method
of
Estimation
Tumor
Site
Species
Sex
Slope
Factor
(
mg/
kg­
day)­
1
Unit
Risk
(
µ
g/
L)­
1
LED
10
(
µ
g/
kg­
day)
10­
5
Risk
Concentration
(
µ
g/
L)
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
77
LMS
Method
Using
BW3/
4
Conversion
U.
S.
EPA
(
1998b)
Large
intestine
Rat
F
4.6
×
10­
3
1.3
×
10­
7
­
77
U.
S.
EPA
(
1994b)*
Large
intestine
Rat
F
7.9
×
10­
3
2.3
×
10­
7
­
40
LED
10/
Linear
Method
U.
S.
EPA
(
1998b)
Large
intestine
Rat
F
4.5
×
10­
3
1.3x10­
7
2.2
×
104
78
*
Adapted
from
IRIS
(
1993b)
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
78
D.
Summary
Table
VIII­
20
Summary
of
Advisory
Values
for
Bromodichloromethane,
Dibromochloromethane,
and
Bromoform
Advisory
Value
Reference
Bromodichloromethane
One­
day
HA
for
10­
kg
child
1
mg/
L
Narotsky
et
al.
(
1997)

Ten­
day
HA
for
10­
kg
child
0.6
mg/
L
NTP
(
1998)

Longer­
term
HA
for
10­
kg
child
0.6
mg/
L
CCC
(
2000d)

Longer­
term
HA
for
70­
kg
adult
2
mg/
L
CCC
(
2000d)

RfD
0.003
mg/
kg­
day
Aida
et
al.
(
1992b)

DWEL
100
µ
g/
L
Aida
et
al.
(
1992b)

Lifetime
HA
Not
applicable
­­

Oral
Slope
Factor
c
3.5
x
10­
2
(
mg/
kg­
day)­
1
NTP
(
1987)

Concentration
for
excess
cancer
risk
(
10­
6)
1.0
µ
g/
L
NTP
(
1987)

Unit
Risk
1.0
x
10­
6
(
µ
g/
L)­
1
NTP
(
1987)

Dibromochloromethane
One­
day
HA
for
10­
kg
child
b
0.6
mg/
L
Aida
et
al.
(
1992a)

Ten­
day
HA
for
10­
kg
child
0.6
mg/
L
Aida
et
al.
(
1992a)

Longer­
term
HA
for
10­
kg
child
0.2
mg/
L
NTP
(
1985)

Longer­
term
HA
for
70­
kg
adult
0.6
mg/
L
NTP
(
1985)

RfD
0.02
mg/
kg­
day
NTP
(
1985)

DWEL
700
µ
g/
L
NTP
(
1985)

Lifetime
HA
60
µ
g/
L
NTP
(
1985)

Oral
Slope
Factor
c
4.3
×
10­
2
(
mg/
kg­
day)­
1
NTP
(
1985)

Concentration
for
Excess
cancer
risk
(
10­
6)
0.8
µ
g/
L
NTP
(
1985)

Unit
Risk
1.2
x
10­
6
(
µ
g/
L)­
1
NTP
(
1985)

Bromoform
One­
day
HA
for
10­
kg
child
5
mg/
L
Burton­
Fanning
(
1901)

Ten­
day
HA
for
10­
kg
child
0.2
mg/
L
NTP
(
1989a)

Longer­
term
HA
for
10­
kg
child
a
0.2
mg/
L
NTP
(
1989a)

Longer­
term
HA
for
70­
kg
adult
0.9
mg/
L
NTP
(
1989a)

RfD
0.03
mg/
kg­
day
NTP
(
1989a)

DWEL
1000
µ
g/
L
NTP
(
1989a)
Table
VIII­
20
(
cont.)

Advisory
Value
Reference
Draft
­
Do
Not
Cite
or
Quote
February
20,
2003
VIII
­
79
Lifetime
HA
Not
applicable
­­

Oral
Slope
Factor
c
4.56
×
10­
3
(
mg/
kg­
day)­
1
NTP
(
1989a)

Concentration
for
Excess
cancer
risk
(
10­
6)
8
µ
g/
L
NTP
(
1989a)

Unit
Risk
1.3
x
10­
7
(
µ
g/
L)­
1
NTP
(
1989a)

a
The
calculated
value
for
the
Longer­
term
HA
was
slightly
higher
than
the
values
for
the
Ten­
day
HA.
Therefore,
use
of
the
Ten­
day
HA
for
a
10­
kg
child
is
recommended
as
an
estimate
of
the
Longer­
term
HA
for
a
10­
kg
child.
b
Use
of
the
Ten­
day
HA
recommended
as
a
conservative
estimate
of
the
One­
day
HA
for
a
10­
kg
child.
c
Use
of
the
Longer­
term
HA
recommended
as
a
conservative
estimate
of
the
Ten­
day
HA
for
a
10­
kg
child.
d
The
oral
slope
factor
was
calculated
using
the
Linearized
Multistage
model
and
an
animal­
to­
human
scaling
factor
of
body
weight3/
4
Abbreviations:
BW
=
Body
weight;
DWEL
=
Drinking
water
exposure
limit;
HA
=
Health
advisory;
LMS
=
Linearized
Multistage
Model
Draft
­
Do
not
cite
or
quote
February
20,
2002
IX
­
1
IX.
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W.
Mello
and
R.
F.
Thomas.
1983.
The
ground
water
supply
survey
summary
of
volatile
organic
contaminant
occurrence
data.
Cincinnati,
OH:
U.
S.
Environmental
Protection
Agency,
Technical
Support
Division,
Office
of
Drinking
Water
and
Office
of
Water.
(
As
cited
in
U.
S.
EPA,
1994b.)

Windham,
G.
C.,
K.
Waller,
M.
Anderson,
L.
Fenster,
P.
Mendola,
and
S.
Swan.
2003.
Chlorination
by­
products
in
drinking
water
and
menstrual
cycle
function.
Env.
Health
Perspect.
In
press.

Windholz,
M.,
ed.
1976.
The
Merck
Index.
Rathway,
NJ:
Merck
&
Co.,
Inc.
(
As
cited
in
U.
S.
EPA,
1994b).

Wolf,
C.
R.,
D.
Mansuy,
W.
Nastainczyk,
G.
Deutschmann
and
V.
Ullrich.
1977.
The
reduction
of
polyhalogenated
methanes
by
liver
microsomal
cytochrome
P­
450.
Mol.
Pharmacol.
13:
698­
705.
(
As
cited
in
U.
S.
EPA,
1994b.)

Yang,
C­
Y.,
B­
H.
Cheng,
S­
S.
Tsai,
et
al.
2000.
Association
between
chlorination
of
drinking
water
and
adverse
pregnancy
outcome
in
Taiwan.
Environ.
Health
Perspect.
108:
765­
768.

Young,
T.
B.,
M.
S.
Kanarek
and
A.
A.
Tslatis.
1981.
Epidemiologic
study
of
drinking
water
chlorination
and
Wisconsin
female
cancer
mortality.
J.
Natl.
Cancer
Inst.
67:
1191­
1198.
(
As
cited
in
U.
S.
EPA,
1994b.)

Young,
T.
B.,
D.
A.
Wolf
and
M.
S.
Kanarek.
1987.
Case­
control
study
of
colon
cancer
and
drinking
water
trihalomethanes
in
Wisconsin.
Int.
J.
Epidemiol.
16(
2):
190­
197.
(
As
cited
in
U.
S.
EPA,
1994b.)

Zeiger,
E.
1990.
Mutagenicity
of
42
chemicals
in
Salmonella.
Environ.
Mol.
Mutagen.
Suppl.
16(
Supplement
18):
32­
54.
(
As
cited
in
U.
S.
EPA,
1994b.)

Zhao,
G.,
amd
J.
W.
Allis.
2002.
Kinetics
of
bromodichloromethane
metabolism
by
cytochrome
P450
isoenzymes
in
human
liver
microsomes.
Chem.
Biol.
Interact.
140:
155­
168.
Draft
­
Do
not
cite
or
quote
February
20,
2002
A
­
1
APPENDIX
A
BENCHMARK
DOSE
MODELING
OF
HEALTH
EFFECTS
ENDPOINTS
FOR
THE
BROMINATED
TRIHALOMETHANES:
BROMODICHLOROMETHANE,
DIBROMOCHLOROMETHANE,
AND
BROMOFORM
A
­
2
Draft
­
Do
not
cite
or
quote
February
20,
2002
A.
INTRODUCTION
The
limitations
of
the
NOAEL/
LOAEL
approach
as
the
basis
for
estimating
thresholds
of
toxic
effect
are
well­
documented
(
e.
g.,
U.
S.
EPA,
1995,
2001b).
These
limitations
include:

1)
the
slope
of
the
dose­
response
plays
little
role
in
determining
the
NOAEL;

2)
the
NOAEL
(
or
LOAEL)
is
limited
to
the
doses
tested
experimentally;

3)
the
determination
of
the
NOAEL
is
based
on
scientific
judgement,
and
is
subject
to
inconsistency;

4)
experiments
using
fewer
animals
tend
to
produce
larger
NOAELs,
and
as
a
result
may
produce
larger
health
advisories
(
HAs)
or
reference
doses
(
RfDs)
(
U.
S.
EPA,
1995,
2001b)
that
may
not
be
sufficiently
protective
of
human
health.

In
contrast,
benchmark
doses
(
BMDs)
are
not
limited
to
the
experimental
doses,
appropriately
reflect
the
sample
size,
and
can
be
defined
in
a
statistically
consistent
manner.
In
light
of
these
considerations,
it
is
becoming
common
practice
to
conduct
assessments
by
performing
BMD
modeling
for
key
endpoints,
in
addition
to
identification
of
NOAELs
and
LOAELs.

This
document
describes
the
analysis
of
the
data
relevant
to
the
development
of
the
Oneday
Ten­
day,
and
Longer­
term
Health
Advisories
(
HAs)
for
bromodichloromethane
(
BDCM),
dibromochloromethane
(
DBCM),
and
bromoform.
Available
data
of
appropriate
duration
were
analyzed
and
the
implications
of
the
calculated
benchmark
doses
for
the
development
of
HAs
were
considered.
Comparisons
of
the
resulting
health
advisories
with
existing
values
are
also
made.
Developmental
and
reproductive
toxicity
studies
were
also
considered
when
effects
were
seen
at
doses
comparable
to
or
lower
than
those
causing
systemic
toxicity
in
subchronic
or
chronic
studies.
The
data
modeled
in
support
of
HA
development
were
also
used
in
derviation
of
the
reference
doses
for
the
three
brominated
trihalomethanes.

B.
SELECTION
OF
STUDIES
AND
ENDPOINTS
FOR
MODELING
The
large
number
of
candidate
data
sets
for
BMD
modeling
­
required
development
of
a
data
prioritization
system.
The
available
studies
were
first
reviewed
for
endpoints
and
data
sets
appropriate
for
BMD
modeling.
Priority
for
modeling
was
given
to
those
endpoints
that
showed
the
greatest
toxicological
relevance
(
e.
g.,
developmental/
reproductive
endpoints,
target
organ
histopathology)
and
ease
of
interpretation.
Ease
of
interpretation
refers
to
the
ability
to
characterize
the
response
as
adverse
and
to
translate
this
into
an
appropriate
response
level
for
input
into
the
BMDS
program.
In
addition,
endpoints
for
which
the
LOAEL
was
less
than
ten
times
the
lowest
LOAEL
observed
in
the
category
(
e.
g.,
1­
day,
10­
day,
and
Longer
term
HAs,
RfD
etc.)
were
given
priority
for
modeling.
Data
considered
for
BMD
and
general
criteria
for
selection
are
listed
in
Tables
A­
1,
A­
2,
and
A­
3.
A
­
3
Draft
­
Do
not
cite
or
quote
February
20,
2002
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
Candidate
Studies
for
Derivation
of
the
One­
day
HA
Lilly
et
al.

(
1994)
Rat
M
Gavage
(
oil)
6
0
200
400
Single
Dose
Kidney
wt
200
400
Yes
Low
No
Model
data
for
aqueous
vehicle
from
this
study
(
see
below)

Rel.
kidney
wt.
200
400
Yes
Low
No
Serum
&
urine
chem
:

Serum
AST,

LDH,
ALT,

Creatinine,
BUN
Urine
pH,
osmolality
200
400
Yes
Generally
low
No
Kidney
histopath
minimal
renal
tubule
degeneration
and
necrosis
­­
200
Yes
High
No
Liver
histopath
minimal
vacuolar
degeneration
200
400
Yes
High
No
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
4
Draft
­
Do
not
cite
or
quote
February
20,
2002
Lilly
et
al.

(
1994)
Rat
M
Gavage
(
aqueous)
6
0
200
400
Single
Dose
Body
wt.
200
400
Yes
Moderate
No
Model
midzonal
vacuolar
degeneration
(
48
hr)
and
renal
tubule
degeneration
(
48
hr)
Table
2,
p.
135
Model
SDH
as
test
Liver
wt
 
200
Yes
Low
No
Rel.
liver
wt.
 
200
Yes
Low
No
Kidney
weights
200
400
Yes
Low
No
Rel
kidney
wt
 
400
NS
Yes
Low
No
Serum
&
urine
chem
Serum
AST,

LDH,
ALT,

Creatinine,

BUN,
urine
pH,

osmolality
200
400
Yes
Generally
low
No
Kidney
histopath
­­
200
Yes
High
Yes
Liver
histopath
200
400
Yes
High
Yes
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
5
Draft
­
Do
not
cite
or
quote
February
20,
2002
Lilly
et
al.

(
1997)
Rat
M
Gavage
(
aqueous)
5
0
123
164
246
328
492
Single
Dose
Body
wt
48
hr
post
328
492
(
for

10%)
Yes
Low
No
Model
SDH
24
hr
post
as
test
Liver
wt
48
hr
post
164
246
Yes
Low
No
Rel
liver
wt
48
hr
post
328
492
Yes
Low
No
Kidney
wt
24
hr
post
246
328
Yes
Low
No
Rel
kidney
24
hr
post
164
246
Yes
Low
No
Rel
kidney
48
hr
post
328
492
Yes
Low
No
Serum
/
urine
chemist
SDH
24
hr
post
­­
123
Yes
Moderate
No
Keegan
et
al.

(
1998)
Rat
Gavage
(
aqueous)
0
21
31
41
82
123
164
246
Single
Dose
Body
wt
82
123
Yes
Moderate
No
Test
model
SDH­

24
hrs
control
group
means
in
Table
2
as
test
Liver
wt.
41
82
Yes
Low
No
Rel
kidney
wt
123
164
Yes
Low
No
Serum
chemistry
(
elevated
ALT,

AST,
and
SDH
activities)
41
82
Yes
Moderate
No
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
6
Draft
­
Do
not
cite
or
quote
February
20,
2002
French
et
al.

(
1999)
Rat
F
Gavage
(
aqueous)
3­
6
0
(
water)

0
(
emulph)

75
150
300
5
days
Body
wt.
150
300
(
FEL)
Yes
Low
(
at
FEL)
No
Most
effects
occurred
at
FEL
PHA
is
sig.
effect
at
dose
below
FEL.
however,

administration
of
vehicle
alone
caused
significant
increase
in
this
endpoint
Spleen
wt.
150
300
Yes
Low
No
Thymus
wt.
150
300
Yes
Low
No
Rel.
thymus
wt.
150
300
Yes
Low
No
MLNC
prolif
ConA
150
300
Yes
Low
No
MLNC
prolif
PHA
75
150
Yes
?
No
Thornton­

Manning
et
al.

(
1994)
Rat
F
Gavage
(
aqueous)
4­
6
0
75
150
300
5
days
Body
wt
150
300
Yes
Moderate
No
Model
liver
histo
path
(
centrilobul.

vacuolar
degener.)

Table
4;
p.
11
Model
kidney
histopath
(
renal
tubule
vacuolar
degeneration
and
regeneration)

Table
5,
p.
13
model
SDH
as
test
Liver
wt
75
150
Yes
Low
No
Rel.
liver
wt.
75
150
Yes
Low
No
Kidney
wt
75
150
Yes
Low
No
Rel.
kidney
wt.
75
150
Yes
Low
No
Serum
chemistry
(
hepatotoxicity)
75
150
Yes
Moderate
No
Serum
chemistry
(
renal
toxicity)
150
300
Yes
Low
No
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
7
Draft
­
Do
not
cite
or
quote
February
20,
2002
Liver/
kidney
histopath
(
mild
to
moderate
centrilobular
hepatocell.

vacuolar
degeneration,

mild
renal
tubule
vacuolar
degener.)
75
150
Yes
High
Yes
Thornton­

Manning
et
al.

(
1994)
Mouse
F
Gavage
(
aqueous)
5­
6
0
75
150
5
days
Liver
wt
75
150
Yes
Low
No
Serum
chemistry
SDH
­­
75
Yes
Moderate
No
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
8
Draft
­
Do
not
cite
or
quote
February
20,
2002
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
Aida
et
al.

(
1992a)
Rat
M
F
Feed
7
7
0
21
62
189
M
204
F
1
month
Body
wt
(
M)
62
189
Yes
Moderate
No
Model
Liver
cell
vacuol.
in
females
Table
8
p.
129
Liver
wt
(
M)
62
189
Yes
Low
No
Kidney
wt
(
M)
62
189
Yes
Low
No
Body
wt.
(
F)
62
204
Yes
Moderate
No
Rel.
liver
wt.
(
F)
62
204
Yes
Low
No
Liver
histopath
(
M)
62
189
Yes
High
No
Liver
histopath
(
F)
21
62
Yes
High
Yes
Chu
et
al.

(
1982a)
Rat
M
Drinking
water
10
0
0.8
8.0
68
28
days
Clinical
signs
serum
chemistry
histology
68
­­
Yes
Low
No
Lack
of
effect;
No
data
selected
for
modelling
Condie
et
al.

(
1983)
Mouse
Gavage
(
oil)
9­
10
0
37
74
148
14
days
Serum
enzymes
[
elevated
SPGT/
ALT]
74
148
Yes
Moderate
No
Model
kidney
histopath
(
Epithelial
hyperplas.);

Table
4,
p.
571
Liver
histopath
(
centrilob.
pallor)

Table
5,
p.
572
Decreased
PAH
uptake
in
vitro
37
74
Yes
Low
No
Liver
Histopath
37?
74
Yes
Mod.
­
High
Yes
Kidney
Histopath
74
148
Yes
High
Yes
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
9
Draft
­
Do
not
cite
or
quote
February
20,
2002
Melnick
et
al.

(
1998)
Mouse
Gavage
(
oil)
10
0
75/
54
150/
107
326/
233
3
weeks
(
5
d/
wk)
Liver
wt
75
150
Yes
Low
No
Model
hepatocyte
hydropic
degener.

Fig.
4,
p.
142
Serum
chem
­
75
Yes
Moderate
No
Liver
histopath
75
150
Yes
High
Yes
Labeling
index
75
150
Yes
Moderate
No
Munson
et
al.

(
1982)
Mouse
M
F
Gavage
(
aq)
8­
12
0
50
125
250
14
days
Body
wt
(
M)
125
250
Yes
Moderate
No
Rel
liver
wt
(
M)
50
125
Yes
Low
No
Spleen
wt
(
M)
125
250
Yes
Low
No
Serum
chem
(
M)

SPGT
SGOT
125
250
Yes
Low
­
moderate
No
Hematology
(
M)
125
250
Yes
Low
No
Hemagglut
(
M)
50
125
Yes
Low
No
Body
wt
(
F)
125
250
Yes
Moderate
No
Rel
liver
wt
(
F)
50
125
Yes
Low
No
Spleen
wt
(
F)
50
125
Yes
Low
No
Rel
Spleen
wt
(
F)
50
125
Yes
Low
No
Serum
chem
(
F)

SPGT
SGOT
125
250
yes
Low
No
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
10
Draft
­
Do
not
cite
or
quote
February
20,
2002
Hematology
(
F)
125
250
Yes
Low
No
AFC/
spleen
(
F)
50
125
Yes
Low­
Moderate?
No
Hemagglutin
(
F)
125
250
Yes
Low
No
NTP
(
1987)
Rat
M
F
Gavage
(
oil)
5M
4­
5F
0
38
75
150
300
600
14
days
Body
wt
(
M)
150
300
Yes
Moderate
No
LOAEL
for
effect
higher
than
other
for
other
endpoints
NTP
(
1987)
Mouse
M
F
Gavage
(
oil)
5M
4­
5F
0
19
38
75
150
300
14
days
Mortality,

lethargy,
gross
renal
pathology
75
150
(
FEL)
Yes
Low
No
No
suitably
sensitive
endpoint
NTP
(
1998)
Rat
M
F
Drinking
water
6
0
9
38
67
(
Grp
A
males)
30
days
Liver
histopath
9
38
Yes
High
Yes
Hepatocyte
indiv.

cell
necrosis
Table
2,
p.
36
Narotsky
et
al.

(
1997)
*
Rat
F
Gavage
(
oil)

(
aq.)
12­

14
0
25
50
75
Gestation
days
6­
15
Full­
litter
resorption
25
50
Yes
High
Yes
Model
full
litter
resorption
(
aqueous
vehicle)

Fig.
2,
incidence
reported
in
text
above
table
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
11
Draft
­
Do
not
cite
or
quote
February
20,
2002
Bielmeier
et
al.

(
2001)
Rat
F
Gavage
(
aq.)
8­
11
0
75
100
Gestation
day
9
Full­
litter
resorption
­­
75
Yes
High
Yes
Model
full
litter
resorption
Table
p.
23
of
manuscript
("
Hormone
profile
II")

Coffin
et
al.
(
2000)
Mouse
Gavage
(
corn
oil)
10
0
150
300
11
days
Liver
histopathology
­
150
Yes
High
No
Aida
et
al.
(
1992a)

used
an
additional,

lower
dose
which
provides
more
information
about
shape
of
curve
in
low
dose
region
for
histopath.

effects.
No
incidence
data
for
histopathology.

Increased
labeling
index
­
200
Yes
Moderate
No
Increased
relative
liver
wt.
­
200
Yes
Low
No
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
NTP
(
1987)
Rat
M
F
Gavage
(
oil)
9­
10
0/
0
19/
14
38/
27
75/
54
150/
107
300/
214
13
weeks
(
5
d/
wk)
Body
weight
75/
54
(
M)

150/
107
(
F)
150/
107
(
M)
300/

214
(
F)
Yes
Moderate
No
Data
in
text
on
p.

35­
36
Histopath
effects
occurred
only
at
FEL
Hepatic
and
renal
histopath
(
M)
150/
107
300/

214
Yes
High
No
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
12
Draft
­
Do
not
cite
or
quote
February
20,
2002
NTP
(
1987)
Mouse
M
Gavage
(
oil)
10
0
6.25/
4.5
12.5/
9
25/
18
50/
36
100/
71
13
weeks
(
5
d/
wk)
Renal
histopath
50/
36
100/
71
Yes
High
Yes
BMD
modelling
conducted
by
ICF
on
data
for
focal
necrosis
of
renal
tubular
epithelium
in
males
NTP
(
1987)
Mouse
F
Gavage
(
oil)
10
0
25/
18
50/
36
100/
71
200/
142
400/
284
13
weeks
(
5
d/
wk)
Liver
histopath
Vacuolated
cytoplasm
50/
36
100/
71
Yes
High
Yes
Data
in
text
on
p.

49
Model
vacuolated
cytoplasm
Reproductive
and
Developmental
Studies
Ruddick
et
al.

(
1983)
Rat
Gavage
(
oil)
9­
14
0
50
100
200
GD
6­
15
Sternebral
aberrations
100
200
Yes
High
Yes
Model
sternebra
variations
Narotsky
et
al.

(
1997)
Rat
Gavage
(
oil)
12­

14
0
25
50
75
GD
6­
15
Developmental
Full
litter
resorption
25
50
Yes
Moderate
(
vehicle)
No
Model
aq.
data.

from
same
study
Narotsky
et
al.

(
1997)
Rat
Gavage
(
aq)
12­

14
0
25
50
75
GD
6­
15
Developmental
Full
litter
resorption
25
50
Yes
High
Yes
This
study
listed
in
table
for
Longer
term
HA.

Maternal
Reduced
body
weight
gain
­
25
Yes
Moderate
to
high
Yes
CCC
(
2000a)
Rabbit
Drinking
water
5
0
4.9
13.9
32.3
76.3
Gestation
days
6­
29
Reproductive
developmental
endpoints
76
­­
No
Potentially
High
No
No
adverse
effects;
small
sample
size
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
13
Draft
­
Do
not
cite
or
quote
February
20,
2002
CCC
(
2000b)
Rabbit
Drinking
water
25
0
1.4
13.4
35.6
55.3
Gestation
days
6­
29
Reproductive/
developmental
55
­­
­­
Potentially
High
No
No
adverse
repro.

or
develop.

effects.
Model
corrected
maternal
wt.
gain
gd
6­
29
as
maternal
effect.

Maternal
Reduced
body
weight
gain
13.4
36
Yes
Moderate
to
high
Yes
CCC
(
2000c)
Rat
Drinking
water
10
Females
0
ppm
50
ppm
150
ppm
450
ppm
1350
ppm
Gestation
days
0­
21
Reproductive/
developmental
50
ppm
150
­­
Potentially
High
No
Decreased
pup
wt.

and
wt.
gain
at
doses
that
caused
parental
toxicity;

reliable
mg/

kgday
dose
could
not
be
estimated.

CCC
(
2000d)
Rat
Drinking
water
25
0.0
2.2
18.4
45.0
82.0
Gestation
days
6­
21
Developmental
Reduced
number
of
ossification
sites
in
phalanges
and
metatarsals
45
82
Yes
Moderate
No
Reversible
variation
occurring
at
doses
that
cause
maternal
toxicity
Maternal
Reduced
body
weight
gain
18.4
45
Yes
Moderate
to
high
Yes
Model
body
weight
gain
for
gestation
days
6­
7
and
6­
9.
Table
A­
1
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromodichloromethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(
mg/
kgday
LOAEL
(
mg/
kgday
L<
10*
LL*
Ease
of
Interp/
Toxi
cologic
Relevance
BMD
model
?
Comments
A
­
14
Draft
­
Do
not
cite
or
quote
February
20,
2002
Bielmeier
et
al.

(
2001)
Rat
Gavage
(
aq.)
8­
11
0
75
100
Gestation
day
9
Full­
litter
resorption
­­
75
Yes
High
Yes
Model
full
litter
resorption
Table
p.
23
of
manuscript
("
Hormone
profile
II")
Study
also
listed
under
Longerterm
HA
*
L<
LL*
10
:
LOAEL
for
endpoint
less
than
10
times
the
lowest
LOAEL
observed
across
all
studies
in
category
A
­
15
Draft
­
Do
not
cite
or
quote
February
20,
2002
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dibromochloromethane
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
Candidate
Studies
for
Derivation
of
the
One­
day
HA
­
none
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
Aida
et
al.

(
1992a)
Rat
M
F
Feed
7
Males
0
18
56
173
Females
0
34
101
332
1
month
Liver
wt
(
M)
56
173
Yes
Moderate
No
Model:

liver
histopath
(
liver
cell
vacuolization
in
M
and
F
Table
8,

p.
129
Rel
liver
wt
(
M)
56
173
Yes
Moderate
No
Liver
histopath
(
M)
56
173
Yes
High
Yes
Body
wt
(
F)
101
332
Yes
Moderate
No
Liver
wt
(
F)
34
101
Yes
Low
No
Rel
liver
wt
(
F)
­­
34
Yes
Low
No
Rel
kidney
wt
(
F)
101
332
Yes
Low
No
Liver
histopath
(
F)
101
332
Yes
High
Yes
Chu
et
al.

(
1982a)
Rat
M
Drinking
water
10
0
.7
8.5
68
28
days
­­
68
­­
­­
­­
­­
No
adverse
effects
observed
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
16
Draft
­
Do
not
cite
or
quote
February
20,
2002
Condie
et
al.

(
1983)
Mouse
M
Gavage
(
oil)
8­
16
0
37
74
147
14
days
Serum
SPGT
74
147
Yes
Moderate
No
Model
renal
mesangial
hypertrophy
and
hepatic
cytoplasmic
vacuolation
Table
5,

p.
572
Liver
histopath
74
147
Yes
High
Yes
Renal
histopath
74
147
Yes
High
Yes
Melnick
et
al.

(
1998)
Mouse
F
Gavage
(
oil)
10
0
50/
37
100/
71
192/
137
417/
298
3
weeks
(
5
d/
wk)
Relative
Liver
wt
­­
50
Yes
Low
No
Model
Incidence
of
hepatocyte
hydropic
degeneration
Table
4,

p.
142
Serum
ALT
100
192
Yes
Moderate
No
Serum
SDH
­­
50
Yes
Moderate
No
Liver
histopath
100
192
Yes
High
Yes
Inc.
labeling
index
192
417
Yes
High
No
Munson
et
al.

(
1982)
Mouse
M
F
Gavage
(
aq)
8­
12
0
50
125
250
14
days
Body
wt.
(
M)
125
150
Yes
Moderate
No
Rel
liver
wt
(
M)
50
125
Yes
Low
No
Spleen
wt
(
M)
125
250
Yes
Moderate?
No
Rel
spleen
wt
(
M)
125
250
Yes
Moderate?
No
Hematology
­

Fibr
(
M)
125
250
Yes
Low
No
Serum
chem
SGPT
(
M)
125
250
Yes
Low
No
AFC/
Spleen
(
M)
125
250
Yes
Moderate
No
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
17
Draft
­
Do
not
cite
or
quote
February
20,
2002
*
AFC/
106
(
M)
50
125
Yes
Moderate
No
Liver
wt
(
F)
125
250
Yes
Low
No
Rel
liver
wt
(
F)
50
125
Yes
Low
No
Hematology
­

Fibr
(
F)
125
250
Yes
Low
No
Serum
SGPT
(
F)
125
250
Yes
?
No
AFC/
spleen
(
F)
125
250
Yes
Moderate
No
AFC/
106
(
F)
50
125
Yes
Moderate
No
NTP
(
1985)
Rat
M
F
Gavage
(
oil)
5
0
60
125
250
500
1000
14
days
Body
wt
(
M)
250
500
(
FEL)
No
Moderate
No
Tables
3
and
4
p.
33
Effects
observed
only
at
levels
where
reduced
survival
occurred:

survival(
2/
5
and
0/
5
for
M
and
F,
respectively
at
the
250
mg/
kg­
day
LOAEL
Dark'd
kid
medulla
(
M)
250
500
No
Low
No
Mottled
liver
(
M)
500
1000
No
Low
No
Dark'd
kid
medulla
(
F)
250
500
No
Low
No
Mottled
liver
(
F)
250
500
No
Low
No
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
18
Draft
­
Do
not
cite
or
quote
February
20,
2002
NTP
(
1985)
Mouse
M
F
Gavage
(
oil)
5
0
30
60
125
250
500
14
days
Stomach
nodules
(
F)
125
250
Yes
Moderate
No
Model
stomach
nodules
in
M
and
F
Table
13
p.
44
Renal
and
hepatic
effects
observed
only
at
levels
where
reduced
survival
observed:

Survival
at
500
mg/

kgday
1/
5
and
2/
5
for
M
and
F
respectively
Stomach
nodules
(
M)
60
125
Yes
Moderate
Yes
Red'd
kid
medulla
(
F)
250
500
No
Low
No
Red'd
kid
medulla
(
M)
250
500
No
Low
No
Mottled
liver
(
M)
125
250
Yes
Low
No
Mottled
liver
(
F)
125
250
Yes
Low
No
Coffin
et
al.
(
2000)
Mouse
F
Gavage
(
oil)
10
0
100
300
11
days
Liver
histopathology
­
100
Yes
High
No
Other
studies
showing
histopath.

effects
used
lower
range
of
doses.
No
incidence
data
for
histopathology

Increased
labeling
index
­
100
Yes
Moderate
No
Increased
relative
liver
wt.
­
100
Yes
Low
No
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
19
Draft
­
Do
not
cite
or
quote
February
20,
2002
NTP
(
1996)
Rat
Drinking
water
10
Males
0
4.2
12.4
28.2
Group
A
Females
0
6.3
17.4
46.0
Group
B
Females
0
7.1
20
47.8
29
days
No
clearly
treatmentrelated
adverse
effects
observed
28
­­
­­
­­
­­
Decreased
wt
gain
observed
in
some
groups,
but
effect
did
not
reach
statistical
significance
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
20
Draft
­
Do
not
cite
or
quote
February
20,
2002
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
Chu
et
al.

(
1982b)
Rat
M
F
Drinking
water
20
Males
0
0.57
6.1
49
224
Females
0
0.64
6.9
55
236
90
days
Liver
histopath
­

prevalence
(
M)
?
?
Yes
High
Yes
Model
incidence
data
Tables
5
and
6
"
treatment"

results
Liver
histopath
­

prevalence
(
M)
49
224
Yes
High
Yes
Daniel
et
al.

(
1990)
Rat
M
F
Gavage
(
oil)
10
0
50
100
200
90
days
Hepatic
and
renal
lesions
Modeled
previously
by
ICF
­­
50
Yes
High
­
Crump
BMDL
10
=

4.2
(
kidney
cortex
degeneration
in
females)

NTP
(
1985)
Rat
Gavage
(
oil)
10
0
15
30
60
125
250
13
weeks
(
5
d/
wk)
hepatic
lesions
Modeled
previously
by
ICF
30
60
Yes
High
­
Crump
BMDL
10
=

0.93
(
liver
fatty
metamorpho
sis
in
males)
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
21
Draft
­
Do
not
cite
or
quote
February
20,
2002
NTP
(
1985)
Mouse
M
F
Gavage
(
oil)
10
0
15
30
60
125
250
13
weeks
(
5
d/
wk)
Liver
histopath
(
M)
Kidney
histopath
(
F)
125
250
Yes
High
No
Occurred
only
at
highest
dose;

data
provided
only
for
0,

125,
and
250
mg/
kg
doses;
Incidence
at
125
0/
10
for
all
endpoints;

5/
10
for
hepatic
vac.

change
and
nephropathy
in
males;
incidence
at
15,
30,
and
60
not
examined.
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
22
Draft
­
Do
not
cite
or
quote
February
20,
2002
Reproductive
and
Developmental
Studies
Borzelleca
and
Carchman
(
1982)
Mouse
M
F
Drinking
Water
10
M
30
F
0
17
171
685
25­
27
weeks
Postnatal
body
wt.

(
cannot
be
modeled
due
to
insufficient
data
on
number
of
litters
evaluated)
­
17
(
marginal
)
Yes
High
No
Marginal
LOAEL
for
parental
toxicity
is
17
mg/
kg­
day
Ruddick
et
al.

(
1983)
Rat
F
Gavage
(
Corn
oil)
9­
14
0
50
100
200
g.
d.
6­
15
­
None
identified
None
identified
­­
­­
No
No
clearly
adverse
effect
NTP
(
1996)
Rat
M
Drinking
Water
10
4.2
12.4
28.2
29
days
­
28.2
­­
­­
­­
No
No
clearly
adverse
effect
on
any
reproductive
endpoint
at
tested
doses
NTP
(
1996)
Rat
F
Drinking
Water
10
6.3
17.4
46.0
35
days
­
46.0
­­
­­
­­
No
No
clearly
adverse
effect
on
any
reproductive
or
development
al
endpoint
at
tested
doses
NTP
(
1996)
Rat
F
Drinking
Water
7.1
20.0
47.8
13
6
days
­
47.8
­­
­­
­­
­
No
clearly
adverse
effect
on
any
reprod
or
develop
endpoint
at
tested
doses
Table
A­
2
Candidate
Studies
and
Data
for
BMD
Modeling
­
Dichlorobromomethane
(
cont.)

A
­
23
Draft
­
Do
not
cite
or
quote
February
20,
2002
*
L<
LL*
10
:
LOAEL
for
endpoint
less
than
10
times
the
lowest
LOAEL
observed
across
all
studies
in
category
FEL,
Frank
effect
level
A
­
24
Draft
­
Do
not
cite
or
quote
February
20,
2002
Table
A­
3
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromoform
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
Candidate
Studies
for
Derivation
of
the
One­
day
HA
for
Bromoform
­
No
suitable
studies
Candidate
Studies
for
Derivation
of
the
Ten­
day
HA
for
Bromoform
Aida
et
al.

(
1992a)
Rat
M
F
Feed
7
Males
0
62
187
618
Females
0
56
208
728
1
month
Liver
histopath
(
M)
62
187
Yes
High
Yes
Model
liver
cell
vacuolization
in
M
and
F
Table
7,
p.
128
Serum
LDH
56
208
Yes
Low
No
BUN
(
F)
56
208
Yes
Low
No
Liver
histopath.

(
F)
56
208
Yes
High
Yes
Chu
et
al.

(
1982a)
Rat
M
Drinking
water
20
0
0.7
8.5
80
28
days
None
80
­­
Yes
­­
No
No
adverse
effects
Condie
et
al.

(
1983)
Mouse
M
Gavage
(
oil)
5­
16
0
72
145
289
14
days
Renal
slice
uptake
PAH
145
289
Yes
Low
No
Model
Liver
histopath:
centrilobular
pallor
Model
kidney
histopath:
mesangial
nephrosis
Table
4,
p.
571
Renal
Histopath
145
289
Yes
High
Yes
Liver
histopath
145
289
Yes
High
Yes
SGPT
activity
145
289
Yes
Low
No
Table
A­
3
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromoform
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
25
Draft
­
Do
not
cite
or
quote
February
20,
2002
Melnick
et
al.

(
1998)
Mouse
F
Gavage
(
oil)
10
0
200
500
3
weeks
(
5
d/
wk)
Rel
Liver
wt
200
500
Yes
Low
No
Stat
sign
increase
in
rel
liver
wt
at
200
­
reported
to
be
about
17%
in
text.
Value
of
200
for
NOAEL
based
on
consistency
of
effects
at
higher
dose
per
Mantus
Model:
liver
hydropic
degeneration
Graph
p.
140
Serum
chemistry
ALT
200
500
Yes
?
No
Serum
Chemistry
SDH
200
500
Yes
?
?

Liver
histopath
200
500
Yes
High
Yes
Labeling
index
200
500
Yes
Moderate
No
Coffin
et
al.
(
2000)
Mouse
F
Gavage
(
oil)
10
0
200
500
11
days
Liver
histopath.
­
200
Yes
High
No
Other
studies
with
lower
range
of
dose.
No
incidence
data
for
histopathology.

Labeling
index
­
200
Yes
Moderate
No
Table
A­
3
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromoform
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
26
Draft
­
Do
not
cite
or
quote
February
20,
2002
Munson
et
al.

(
1982)
Mouse
M
F
Gavage
(
aq.)
7­
12
0
50
125
250
14
days
Liver
wt
(
M)
50
125
Yes
Low
No
Rel
liver
wt
(
M)
50
125
Yes
Low
No
Hematology
­

Fibr
(
M)
125
250
Yes
Low
No
*
Serum
SGOT
(
M)
125
250
Yes
Low
No
NTP
(
1989a)
Mouse
M
F
Gavage
(
oil)
5
Male
0
50
100
200
400
600
Female
0
100
200
400
600
800
14
days
Stomach
nodules
(
M)
200
400
Yes
Moderate
Yes
Model
incidence
of
stomach
nodules
in
males
p.
45
Males:

400
4/
5
600
3/
5
Females:

600
2/
5
800
1/
5
Stomach
nodules
(
F)
400
600
Yes
Moderate
No
NTP
(
1989a)
Rat
M
F
Gavage
(
oil)
5
0
100
200
400
600
800
14
days
Body
wt
(
M)
200
400
Yes
Moderate
No
Possibly
model
body
weight
p.
36
Table
A­
3
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromoform
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
27
Draft
­
Do
not
cite
or
quote
February
20,
2002
Candidate
Studies
for
Derivation
of
the
Longer­
term
HA
Chu
et
al.

(
1982b)
Rat
M
F
Drinkin
g
water
9­
10
Males
0
0.65
6.1
57
218
Female
s
0
0.64
6.9
55
283
90
days
Liver
Histopath
(
M)
57
218
Yes
High
Yes
Incidence
and
mean
severity
score
provided
for
combined
hepatic
lesions.

Model
'
treatment'
prevalence
for
liver
lesions
in
M
and
F
Serum
chem
data
(
LDH)

presented
only
for
high
dose
(
Insuff
data
for
modeling)

Liver
Histopath
(
F)
55
283
Yes
High
Yes
NTP
(
1989a)
Rat
M
F
Gavage
(
corn
oil)
10
0
12
25
50
100
200
13
weeks
(
5
d/
wk)
Liver
Histopath.
25
50
(
hepati
c
vacuoli
zation)
Yes
High
­
Previously
calculated
Crump
BMDL10
2.65
(
hepatic
vacuolization
in
male
rats)
Table
A­
3
Candidate
Studies
and
Data
for
BMD
Modeling
­
Bromoform
(
cont.)

Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<
10*
LL*
Ease
of
Interpretation
Toxicologic
Relevance
BMD
model
?
Comments
A
­
28
Draft
­
Do
not
cite
or
quote
February
20,
2002
NTP
(
1989a)
Mouse
Gavage
(
corn
oil)
10
0
25
50
100
200
400
13
weeks
(
5
d/
wk)
Liver
Histopath
100
200
Yes
High
Yes
Model
hepatic
vacuolization
Reproductive
and
Developmental
Studies
Ruddick
et
al.

(
1983)
F
Gavage
(
Corn
oil)
14­

15
0
50
100
200
gd
6­
15
Sternebra
aberrations
50
100
Yes
High
Yes
Dose­
dependent
increase
in
sternebra
aberrations;
intraparietal
deviations
at
midand
high
doses.

Intraparietal
variations
­
­
­
High
No
NTP
(
1989b)
Mouse
M
F
Gavage
(
oil)
20
20
0
50
100
200
(
NOAE
L)
105
days
No
adverse
effects
at
doses
tested
200
­
­
High
No
No
detectable
effect
on
fertility,

litters/
pair,
live
pups/
litter;

proportion
of
live
births,
sex
of
live
pups,
or
pup
body
weight.

*
L<
LL*
10
:
LOAEL
for
endpoint
less
than
10
times
the
lowest
LOAEL
observed
across
all
studies
in
category
A
­
29
Draft
­
Do
not
cite
or
quote
February
20,
2002
C.
METHODS
Benchmark
Dose
The
brominated
trihalomethane
data
sets
considered
for
dose­
response
modeling
include
both
quantal
and
continuous
endpoints.
EPA's
Benchmark
Dose
Software
(
BMDS)
(
U.
S.
EPA,
2000a)
was
used
to
accomplish
all
of
the
model
fitting
and
estimation
of
the
BMD
and
lower
95%
confidence
limit
(
BMDL).
The
methods
and
models
applied
to
both
quantal
and
continuous
endpoints
are
presented
here.

Quantal
Models
Seven
of
the
nine
quantal
models
implemented
in
the
BMDS
package
were
used
to
represent
the
dose­
response
behavior
of
the
quantal
endpoints.
Specifically,
the
models
used
were
the
gamma
model,
the
logistic
and
log­
logistic
models,
the
probit
and
log­
probit
models,
the
multistage
model,
and
the
Weibull
model.
Two
other
models,
the
linear
and
the
quadratic
models,
were
not
fit
to
the
data
because
they
are
special
cases
of
both
the
multistage
and
the
Weibull
models.
If
the
fitting
of
the
multistage
or
Weibull
models
resulted
in
a
linear
or
a
quadratic
form,
then
those
result
were
used;
otherwise,
the
linear
or
quadratic
models
would
not
provide
a
fit
as
good
as
the
multistage
or
Weibull
model
and
so
were
not
separately
obtained.

The
equations
defining
each
of
these
models
are
presented
here
(
U.
S.
EPA,
2000a).
In
all
of
the
following,
P(
d)
represents
the
probability
of
response
(
i.
e.,
adverse
effect)
following
exposure
to
"
dose"
d.
In
all
of
these
models,
 ,
 ,
and
 
are
model
parameters
estimated
using
maximum
likelihood
techniques,
as
described
below.

Table
A­
4
Model
Equations
used
in
BMD
Calculations
for
Health
Advisories
Model
Equation
Conditions
gamma
P(
d)
=
 
+
(
1
­
 )
·
(
1/
 (
a))
·

t ­
1e­
tdt
0

 
<
1,
 

0,
and
 
>
1.
 (
x)
is
the
gamma
function,
and
the
integral
runs
from
0
to
 d.

logistic
P(
d)
=
[
1
+
exp{­(
 
+
 d)}]­
1
 

0
loglogistic
P(
d)
=
 
+
(
1
­
 )
·
[
1
+
exp{­(
 
+
 ln(
d))}]­
1
The
log­
logistic
model
has
much
the
same
form
as
the
logistic
model
except
when
d
=
0,
in
which
case
P(
d)
=
g.
In
this
case
b

0,
and
for
the
background
parameter
 ,
0

 
<
1.

probit
P(
d)
=
 (
 
+
 d)
 (
x)
is
the
standard
normal
cumulative
distribution
function
and
 

0.
A
­
30
Draft
­
Do
not
cite
or
quote
February
20,
2002
log­
probit
P(
d)
=
 
+
(
1
­
 )
·
 (
 
+
 
·
ln(
d))
The
log­
probit
model
has
a
form
similar
to
the
probit
model
except
when
d
=
0,
in
which
case
P(
d)
=
 .
Here
0

 
<
1,
and
 

1
multistage
model
P(
d)
=
 
+
(
1
­
 )
·
(
1
­
exp{­(
 
1
d
+
 
2
d2
+
...
+
 
n
dn)})
all
the
 
parameters
are
restricted
to
be
nonnegative
and
0

 
<
1.
When
applied
to
the
brominated
trihalomethane
data
sets
in
these
analyses,
the
degree
of
the
multistage
model
(
the
highest
power
on
dose
in
the
above
equation,
n)
was
set
equal
to
one
less
than
the
number
of
dose
groups
in
the
experiment
being
analyzed.

Weibull
model1
P(
d)
=
 
+
(
1
­
 )(
1
­
exp{­
 d })
The
background
parameter
 
is
restricted
to
fall
between
0
(
inclusive)
and
1,
and
 
is
greater
than
or
equal
to
0.
For
these
analyses,
the
parameter
 
is
constrained
to
be
greater
than
or
equal
to
1.1
1The
linear
model
is
a
special
case
of
the
Weibull
model
obtained
by
fixing
the
parameter
 
equal
to
1.
The
quadratic
model
is
a
special
case
of
the
Weibull
model
obtained
by
fixing
the
parameter
 
equal
to
2.

When
fitting
all
of
the
above­
mentioned
quantal
models,
maximum
likelihood
methods
were
used
to
estimate
the
parameters
of
the
models.
That
method
maximizes
the
log­
transformed
likelihood
of
obtaining
the
observed
data,
which
is
(
except
for
an
additive
constant)
given
by
L
=

[
n
i
·
ln{
P(
d
i)}
+
(
N
i
­
n
i)
·
ln{
1
­
P(
d
i)}]

where
the
sum
runs
over
i
from
1
to
k
(
the
number
of
dose
groups),
and
for
group
i,
d
i
is
the
dose
(
exposure
level),
N
i
is
the
number
of
individuals
tested,
and
n
i
is
the
number
of
individuals
responding
(
U.
S.
EPA,
2000a).

Continuous
Models
The
continuous
endpoints
of
interest
with
respect
to
brominated
trihalomethanes
toxicity
were
quantitatively
summarized
by
group
means
and
measures
of
variability
(
standard
errors
or
standard
deviations).
The
models
used
to
represent
the
dose­
response
behavior
of
those
continuous
endpoints
are
those
implemented
in
EPA's
Benchmark
Dose
Software
(
U.
S.
EPA,
A
­
31
Draft
­
Do
not
cite
or
quote
February
20,
2002
2000a).
These
models
were
the
power
model,
the
Hill
model,
and
the
polynomial
model.
These
mathematical
models
fit
to
the
data
are
defined
here.
In
all
cases,
µ
(
d)
indicates
the
mean
of
the
response
variable
following
exposure
to
"
dose"
d.

The
power
model
is
represented
by
the
equation
µ
(
d)
=
 
+
 d 
where
the
parameter
 
is
restricted
to
be
nonnegative.
[
The
linear
model
is
obtained
when
 
is
fixed
at
a
value
of
1.
The
linear
model
was
not
separately
fit
to
the
data;
if
the
result
of
fitting
the
power
model
does
not
result
in
the
linear
form,
 
=
1,
then
the
linear
model
does
not
fit
as
well
as
the
more
general
power
model,
by
definition.]

The
Hill
model
is
given
by
the
following
equation:

µ
(
d)
=
 
+
(
vdn)
/
(
dn
+
kn))

where
the
parameters
n
and
k
are
restricted
to
be
positive.
Because
the
Hill
model
has
four
parameters
to
be
estimated
(
 ,
v,
n,
and
k),
the
power
n
was
fixed
equal
to
1
when
the
model
was
fit
to
data
sets
with
only
three
dose
groups,
so
that
the
number
of
estimated
parameters
did
not
exceed
the
number
of
data
points.

The
polynomial
model
is
defined
as
µ
(
d)
=
 
0
+
 
1
d
+
...
+
 
n
dn
where
the
degree
of
the
polynomial,
n,
was
set
equal
to
one
less
than
the
number
of
dose
groups
in
the
experiment
being
analyzed.
Note
that
U.
S.
EPA
(
2000a)
recommends
the
use
of
the
most
parsimonious
model
that
provides
an
adequate
fit
to
the
data.
It
may
appear
that
use
of
a
polynomial
model
with
degree
equal
to
one
less
than
the
number
of
dose
groups
would
not
yield
the
most
parsimonious
model.
However,
allowing
the
model
to
have
that
degree
is
not
the
same
as
forcing
the
model
to
have
that
degree;
in
the
model
fitting,
if
fewer
parameters
(
e.
g.,
a
lower
degree
polynomial)
is
adequate
and
consistent
with
the
data,
then
the
fitting
will
reflect
that
fact
and
a
more
parsimonious
model
will
be
the
result.
For
these
analyses,
the
values
of
the
 
parameters
allowed
to
be
estimated
were
constrained
to
be
either
all
nonnegative
or
all
nonpositive
(
as
dictated
by
the
data
set
being
modeled,
i.
e.,
nonnegative
if
the
mean
response
increased
with
increasing
dose
or
nonpositive
if
the
mean
response
decreased
with
increasing
dose).

In
the
case
of
continuous
endpoints,
one
must
assume
something
about
the
distribution
of
individual
observations
around
the
dose­
specific
mean
values
defined
by
the
above
models.
The
assumptions
imposed
by
BMDS
were
used
in
this
analysis:
individual
observations
were
assumed
to
vary
normally
around
the
means
with
variances
given
by
the
following
equation:

 
i
2
=
 2
·
[
µ
(
d
i)]
 
A
­
32
Draft
­
Do
not
cite
or
quote
February
20,
2002
where
both
 2
and
 
were
parameters
estimated
by
the
model.

Given
those
assumptions
about
variation
around
the
means,
maximum
likelihood
methods
were
applied
to
estimate
all
of
the
parameters,
where
the
log­
likelihood
to
be
maximized
is
(
except
for
an
additive
constant)
given
by
L
=

[(
N
i/
2)
·
ln(
 
i
2)
+
(
N
i
­
1)
s
i
2/
2 
i
2
+
N
i{
m
i
­
µ
(
d
i)}
2/
2 
i
2]

where
N
i
is
the
number
of
individuals
in
group
i
exposed
to
dose
d
i,
and
m
i
and
s
i
are
the
observed
mean
and
standard
deviation
for
that
group.
The
summation
runs
over
i
from
1
to
k
(
the
number
of
dose
groups).

Goodness
of
Fit
Analyses
For
the
quantal
models,
goodness
of
fit
was
determined
by
the
modeling
software
using
the
chi­
square
test.
This
test
is
based
on
sums
of
squared
differences
between
observed
and
predicted
numbers
of
responders.
The
degrees
of
freedom
for
the
chi­
square
test
statistic
are
equal
to
the
number
of
dose
groups
minus
the
number
of
parameters
fit
by
the
method
of
maximum
likelihood
(
ignoring
those
parameters
that
are
estimated
to
be
equal
to
one
of
the
bounds
defining
their
constraints
­­
see
the
discussion
above
about
constraints
imposed
on
the
model
parameters).
When
the
number
of
parameters
estimated
equals
the
number
of
dose
groups,
there
are
no
degrees
of
freedom
for
a
statistical
evaluation
of
fit.

For
the
continuous
models,
goodness
of
fit
was
determined
based
on
a
likelihood
ratio
statistic.
In
particular,
the
maximized
log­
likelihood
associated
with
the
fitted
model
was
compared
to
the
log­
likelihood
maximized
with
each
dose
group
considered
to
have
a
mean
and
variance
completely
independent
of
the
means
and
variances
of
the
other
dose
groups.
It
is
always
the
case
that
the
latter
log­
likelihood
will
be
at
least
as
great
as
the
model­
associated
loglikelihood
but
if
the
model
does
a
"
reasonable"
job
of
fitting
the
data,
the
difference
between
the
two
log­
likelihoods
will
not
be
too
great.
A
formal
statistical
test
reflecting
this
idea
uses
the
fact
that
twice
the
difference
in
the
log­
likelihoods
is
distributed
as
a
chi­
square
random
variable.
The
degrees
of
freedom
associated
with
that
chi­
squared
test
statistic
are
equal
to
the
difference
between
the
number
of
parameters
fit
by
the
model
(
including
the
parameters
 2
and
 
defining
how
variances
change
as
a
function
of
mean
response
level)
and
twice
the
number
of
dose
groups
(
which
is
equal
to
the
number
of
parameters
estimated
by
the
"
model"
assuming
independence
of
dose
group
means
and
variances).

Visual
fit,
particularly
in
the
low­
dose
region,
was
assessed
for
models
that
had
acceptable
global
goodness­
of­
fit.
Acceptable
global
goodness
of
fit
was
either
a
p­
value
greater
than
or
equal
to
0.1,
or
a
perfect
fit
when
there
were
no
degrees
of
freedom
for
a
statistical
test
of
fit.
Choice
of
0.1
is
consistent
with
current
U.
S.
EPA
guidance
for
BMD
modelling
(
U.
S.
EPA,
2000b).
Local
fit
was
evaluated
visually
on
the
graphic
output,
by
comparing
the
observed
and
estimated
results
at
each
data
point.
A
­
33
Draft
­
Do
not
cite
or
quote
February
20,
2002
Goodness­
of­
fit
statistics
are
not
designed
to
compare
different
models,
particularly
if
the
different
models
have
different
numbers
of
parameters.
Within
a
family
of
models,
adding
parameters
generally
improves
the
fit.
BMDS
reports
the
Akaike
Information
Criterion
(
AIC)
to
aid
in
comparing
the
fit
of
different
models.
The
AIC
is
defined
as
 
2L+
2p,
where
L
is
the
loglikelihood
at
the
maximum
likelihood
estimates
for
the
parameters,
and
p
is
the
number
of
model
parameters
estimated.
When
comparing
the
fit
of
two
or
more
models
to
a
single
data
set,
the
model
with
the
lesser
AIC
was
considered
to
provide
a
superior
fit.

Definition
of
the
BMR
and
Corresponding
BMD
and
BMDL
For
all
of
the
quantal
endpoints
analyzed
here,
the
BMDs
and
BMDLs
were
defined
based
on
BMRs
of
5%
and
10%
extra
risk.
BMDLs
were
defined
as
the
95%
lower
bound
on
the
corresponding
BMD
estimates.
Confidence
bounds
were
calculated
by
BMDS
using
a
likelihood
profile
method.

Although
the
10%
response
level
was
selected
as
the
"
point
of
departure"
for
all
the
quantal
endpoints
analyzed
here,
we
have
chosen
to
follow
standard
practice
and
include
results
for
both
the
5%
and
10%
level
of
response.
In
some
cases
(
see
discussions
below),
a
comparison
of
the
5%
and
10%
results
gives
clues
about
problems
with
some
of
the
models.
In
general,
we
have
included
both
for
completeness,
as
there
is
no
current
consensus
concerning
the
most
appropriate
point
of
departure
except
in
some
particular
cases
(
e.
g.,
use
of
5%
risk
for
developmental
toxicity
tests
where
the
nesting
of
effects
has
been
modeled
using
models
specifically
designed
for
such
experimental
designs).

For
the
continuous
models,
BMDs
were
implicitly
defined
as
follows:


µ
(
BMD)
­
µ
(
0)

=
 
·
 
1
where
 
1
is
the
model­
estimated
standard
deviation
in
the
control
group.
In
other
words,
the
BMR
was
defined
as
a
change
in
mean
corresponding
to
some
multiplicative
factor
of
the
control
group
standard
deviation.

The
value
of
 
used
in
this
analysis
was
1.1.
This
value
was
chosen
based
on
the
work
of
Crump
(
1995),
who
showed
that
this
choice
corresponded
to
an
additional
risk
of
10%
when
the
background
response
rate
was
assumed
to
be
1%,
with
normal
variation
around
the
mean
(
and
constant
standard
deviation).
Although
the
current
analyses
allowed
for
nonconstant
standard
deviations
and
estimated
extra
risk,
while
the
Crump
(
1995)
comparison
was
based
on
constant
standard
deviations
and
additional
risk,
the
values
of
1.1
was
used
for
two
reasons.
First,
the
difference
between
additional
and
extra
risk
is
small
when
the
background
rate
is
1%
or
less,
so
that
the
change
from
additional
to
extra
risk
will
have
minimal
impact
on
the
correspondences
proven
by
Crump
(
1995).
Second,
there
can
be
no
such
generic,
a
priori
correspondences
when
standard
deviations
are
allowed
to
vary
in
a
manner
determined
only
after
the
model
fitting
is
accomplished.
Thus,
to
avoid
data
set­
and
model­
specific
choices
for
 ,
the
correspondences
proven
by
Crump
(
1995)
can
be
used
as
the
best
available,
consistent
definition
of
the
benchmark
A
­
34
Draft
­
Do
not
cite
or
quote
February
20,
2002
response.
The
definition
of
the
BMR
as
a
change
in
mean
of
1.1
times
the
control
standard
deviation
is
very
close
to
the
default
value
of
1
standard
deviation
recommended
by
recent
draft
EPA
guidelines
(
U.
S.
EPA,
2000b).
In
the
following,
BMDs
and
BMDLs
corresponding
to
 
=
1.1
are
denoted
BMD
10
and
BMDL
10,
because
of
the
just­
noted
association
of
that
value
of
 
with
10%
risk.

As
for
the
quantal
models,
for
all
of
the
continuous
models
BMDLs
were
defined
as
the
95%
lower
bound
on
the
corresponding
BMD.
Confidence
intervals
were
calculated
using
a
profile
likelihood
method.

Choice
of
BMDL
The
following
guidance
was
followed
with
regard
to
the
choice
of
the
BMDL
to
use
as
a
point
of
departure
for
calculation
of
a
health
advisory.
This
guidance
is
consistent
with
recommendations
in
U.
S.
EPA
(
2000b).
For
each
endpoint,
the
following
procedure
is
recommended:

1.
Models
with
an
unacceptable
fit
(
including
consideration
of
local
fit
in
the
low­
dose
region)
are
excluded.
Visual
fit,
particularly
in
the
low­
dose
region,
was
assessed
for
models
that
had
acceptable
global
goodness­
of­
fit.

2.
If
the
BMDL
values
for
the
remaining
models
for
a
given
endpoint
are
within
a
factor
of
3,
no
model
dependence
is
assumed,
and
the
models
are
considered
indistinguishable
in
the
context
of
the
precision
of
the
methods.
The
models
are
then
ranked
according
to
the
AIC,
and
the
model
with
the
lowest
AIC
is
chosen
as
the
basis
for
the
BMDL.

3.
If
the
BMDL
values
are
not
within
a
factor
of
3,
some
model
dependence
is
assumed,
and
the
lowest
BMDL
is
selected
as
a
reasonable
conservative
estimate,
unless
it
is
an
outlier
compared
to
the
results
from
all
of
the
other
models.
Note
that
when
outliers
are
removed,
the
remaining
BMDLs
may
then
be
within
a
factor
of
3,
and
so
the
criteria
given
in
item
2.
would
be
applied.

4.
The
BMDL
values
from
all
modeled
endpoints
are
compared,
along
with
any
NOAELs
or
LOAELs
from
data
sets
that
were
not
amenable
to
modeling,
and
the
lowest
NOAEL
or
BMDL
is
chosen.

5.
Models
with
an
unacceptable
fit
(
including
consideration
of
local
fit
in
the
low­
dose
region)
are
excluded.
A
­
35
Draft
­
Do
not
cite
or
quote
February
20,
2002
D.
MODELING
RESULTS
1.
Bromodichloromethane
The
majority
of
endpoints
modeled
consisted
of
dichotomous
data.
BMDS
modeling
results
for
bromodichloromethane
dichotomous
endpoints
are
summarized
in
Table
A­
5
below.
Four
sets
of
continuous
data
were
modeled.
These
results
are
summarized
in
Section
f
below.
Detailed
output
for
each
model
run
is
compiled
in
Appendix
B,
provided
in
electronic
format
on
compact
disk.

a.
Developmental
and
Reproductive
studies
Three
data
sets
for
developmental
or
reproductive
toxicity
were
modeled.
When
the
data
for
full
litter
resorption
(
FLR)
in
rats
reported
by
Bielmeier
et
al.
(
2000)
were
analyzed,
the
BMDL
results
for
the
log­
logistic
model
were
low
relative
to
the
corresponding
BMD
estimates
(
compared
to
the
estimates
obtained
from
the
other
models);
the
results
from
the
log­
logistic
model
might
be
that
is
considered
qualitative
outliers.
The
remaining
values
are
still
not
within
factor
of
3,
indicating
some
model
dependence
of
the
results.
In
any
case,
the
multistage
model
was
chosen
as
it
gave
the
smallest
value
for
the
AIC.

Modeling
of
FLR
data
from
Narotsky
et
al.
(
1997)
also
gave
the
same
type
of
questionable
results
for
the
log­
logistic
model
(
very
low
BMDLs
relative
to
the
BMD).
Here,
as
in
the
case
of
the
Bielmeier
et
al.
(
2000)
modeling,
the
initial
fit
of
the
log­
logistic
model
does
not
appear
to
be
suspect;
the
goodness
of
fit
evaluations
and
visual
examinations
of
the
model
predictions
are
consistent
with
the
data
and
with
the
other
models.
It
appears
that
there
is
some
error
or
problem
with
the
log­
logistic
model
in
the
BMDS
software
that
affects
the
calculation
of
BMDLs
for
some
data
sets.
When
the
log­
logistic
model
was
eliminated
from
consideration,
the
remaining
BMDLs
are
within
a
factor
of
3.
The
log­
probit
model
was
selected
because
it
has
the
lowest
value
for
the
AIC.

Data
from
the
study
by
Ruddick
et
al.
(
1983)
consisted
of
the
count
of
the
numbers
of
litters
that
had
one
or
more
fetuses
with
sternebral
variations.
Although
this
expression
of
the
response
rates
does
not
correspond
directly
to
the
probability
of
a
response
in
the
offspring
of
treated
dams,
it
is
consistent
with
the
full
litter
resorption
results
from
Bielmeier
et
al.
(
2001)
and
Narotsky
et
al.
(
1997)
in
the
sense
that
it
relates
to
effects
recorded
at
the
level
of
the
dam.
The
log­
logistic
and
log­
probit
models
could
not
determine
BMDLs
for
this
data
set.
However,
the
other
models
did
provide
estimates
of
the
BMDLs,
all
of
which
were
within
a
factor
of
three
of
one
another.
The
multistage,
having
the
lowest
AIC
of
all
the
models
was
selected
as
the
basis
for
the
BMDL
estimate
for
this
data
set.

b.
One­
day
Health
Advisory
Four
data
sets
were
modeled
in
support
of
the
One­
day
HA
for
bromodichloromethane.
For
the
Lilly
et
al.
(
1994)
data
on
vacuolar
degeneration
in
male
rats,
the
multistage
model
gave
questionable
results
(
very
high
AIC
and
a
goodness
of
fit
p
value
that
appeared
unrealistically
high
A
­
36
Draft
­
Do
not
cite
or
quote
February
20,
2002
when
the
model
fit
was
examined
visually)
and
was
eliminated
from
consideration.
The
BMDS
software
gave
warnings
on
bound
calculation
for
the
probit
model.
All
of
the
remaining
BMDLs
are
within
factor
of
3,
so
the
log­
probit
model
was
selected
because
it
has
the
lowest
value
for
the
AIC.

When
the
data
from
the
same
study
for
renal
tubular
degeneration
were
modeled,
the
multistage
and
log­
logistic
gave
questionable
results
and
were
eliminated
from
consideration.
The
remaining
BMDLs
are
not
within
a
factor
of
3,
indicating
some
model
dependence
of
the
results.
The
lowest
BMDL
was
thus
selected
as
a
reasonable
conservative
estimate.
The
gamma
or
Weibull
models
predict
the
same
BMDL.
The
gamma
model
was
selected
on
the
basis
of
having
the
lowest
AIC.

It
is
perhaps
informative
to
compare
the
considerations
applied
here,
in
the
case
of
the
renal
endpoint,
to
those
applied
above
to
the
Narotsky
et
al.
(
1997)
modeling
results.
In
the
case
of
the
Narotsky
et
al.
(
1997)
results,
a
single
model
(
log­
logistic)
gave
a
BMDL
10
that
was
about
8­
fold
lower
than
the
corresponding
BMD
10,
whereas
the
other
models
gave
BMDLs
that
were
within
a
factor
of
about
2
of
the
corresponding
BMDs.
The
discrepancy
was
even
greater
at
a
BMR
of
5%,
suggesting
that
a
problem
may
be
associated
with
that
one
model.
In
the
case
of
the
Lilly
et
al.
(
1994)
renal
effect,
after
eliminating
the
obviously
problematic
model
results
(
loglogistic
and
multistage),
the
differences
between
BMDs
and
corresponding
BMDLs
are
present
and
consistent
for
all
the
models
at
both
5%
and
10%
response.
It
is
true
that
some
models
have
a
greater
difference
between
the
BMD
and
the
BMDL
than
do
other
models
and
that
this
model
dependence
is
due,
at
least
in
part,
to
the
fact
that
the
BMD
estimates
fall
below
the
lowest
nonzero
experimental
dose.
But
the
choice
of
the
most
conservative
BMDL
is
intended
to
cover
that
model
dependence:
if
there
is
little
information
in
the
region
of
interest,
so
that
otherwise
reasonable
(
good­
fitting)
models
disagree
as
to
the
BMDL
because
of
differences
in
possible
curve
shapes,
the
most
conservative
choice
is
a
good
one
since
we
can
not
rule
out
the
possibility
that
the
true
curve
shape
is
described
by
the
most
conservative
model.
This
use
of
the
BMD
methodology
and
treatment
of
model­
dependence
is
much
superior
to
the
choice
of
some
other
(
higher)
BMR
to
use
in
cases
where
response
rate
at
the
lowest
nonzero
experimental
dose
is
greater
than
10%.
Such
an
alternative
would
entail
additional
arbitrary
decisions
about
what
the
higher
BMR
should
be
and
how
to
scale
the
results
corresponding
to
that
BMR
so
as
to
be
consistent
with
results
from
studies
in
which
BMDL
10
s
were
estimated.

Renal
and
hepatic
histopathology
data
were
modeled
from
the
study
of
Thornton­
Manning
et
al.
(
1994).
The
data
for
hepatic
centrilobular
vacuolar
degeneration
were
not
satisfactorily
fit
(
all
goodness
of
fit
p
values
were
less
than
0.1)
by
any
model.
These
data
displayed
some
peculiarities:
0%
response
at
the
lowest
nonzero
dose,
100%
response
at
the
next
dose,
and
then
a
drop
to
75%
response
at
the
highest
dose.
Because
of
the
poor
fit,
the
BMDLs
for
this
endpoint
can
not
be
used.
For
renal
data,
the
multistage
model
gave
questionable
results
(
very
high
AIC
and
a
goodness
of
fit
p
value
that
appeared
unrealistically
high
when
the
model
fit
was
examined
visually)
and
was
eliminated
from
consideration.
The
remaining
BMDLs
were
within
a
factor
of
3,
and
the
log­
logistic
model
was
selected
because
it
has
the
smallest
AIC.
A
­
37
Draft
­
Do
not
cite
or
quote
February
20,
2002
c.
Ten­
day
Health
Advisory
Five
data
sets
were
modeled
using
the
BMDS
software
in
support
of
the
Ten­
day
HA
for
bromodichloromethane.
For
the
Condie
et
al.
(
1983)
data
on
liver
histopathology,
the
logistic
and
probit
models
were
eliminated
on
the
basis
of
poor
fit.
The
rest
of
the
BMDLs
were
within
a
factor
of
3,
so
the
log­
logistic
model
was
selected
on
the
basis
of
the
smallest
AIC.
When
renal
histopathology
data
from
the
same
study
were
modeled,
the
logistic
and
probit
models
were
eliminated
for
lack
of
fit.
The
remaining
BMDLs
were
within
a
factor
of
3,
so
the
models
with
the
lowest
AIC
(
Weibull
and
log­
logistic)
were
examined.
The
Weibull
model
results
were
selected
on
the
basis
of
the
smallest
BMDL.
For
the
Aida
et
al.
(
1992a)
data
set
for
liver
cell
vacuolation,
questionable
results
were
obtained
with
the
log­
logistic
model
(
very
low
BMDLs
relative
to
the
BMD).
The
remaining
BMDLs
are
within
a
factor
of
3
and
the
multistage
model
was
selected
because
it
had
the
smallest
AIC
value.
When
the
Melnick
et
al.
(
1998)
data
were
analyzed,
the
multistage
model
was
eliminated
because
it
gave
a
goodness
of
fit
p­
value
that
was
unrealistically
high
when
the
curve
fit
was
evaluated
by
visual
inspection
and
because
the
AIC
was
very
large.
The
BMDL
values
of
the
remaining
models
were
within
a
factor
of
three.
The
result
from
the
Weibull
model
was
selected
on
the
basis
of
having
the
lowest
AIC
value.
When
results
for
the
NTP
(
1998)
study
were
examined,
all
models
gave
acceptable
fit.
Because
all
BMDLs
were
within
a
factor
of
three
of
one
another,
the
log­
logistic
model
was
selected
on
the
basis
of
the
smallest
AIC
value.

d.
Longer­
term
Health
Advisory
A
single
data
set
(
NTP,
1987)
was
analyzed
using
the
BMDS
software
in
support
of
the
Longer­
term
HA.
Results
for
the
logistic
and
probit
models
were
rejected
for
lack
of
fit
(
goodness
of
fit
p
values
less
than
0.1).
The
remaining
p
values
were
within
a
factor
of
3,
so
the
log­
probit
model
was
selected
on
the
basis
of
the
smallest
AIC.

e.
RfD
Several
data
sets
that
had
previously
given
low
BMDL
10
estimates
when
modeled
using
the
Crump
benchmark
dose
software
(
THC
and
THWC
programs;
K.
S.
Crump,
Inc.)
were
reanalyzed
using
the
BMDS
software
and
current
guidance
for
evaluation
of
results.
An
advantage
of
the
BMDS
package
is
that
it
includes
several
additional
model
options
for
data
analysis.
Results
for
the
models
(
Weibull,
multistage)
common
to
both
programs
were
in
close
agreement.
However,
one
or
more
of
the
additional
models
available
in
the
BMDS
sometimes
fit
the
data
better
when
analyzed
by
the
criteria
set
forth
in
Section
C
(
above).
In
these
cases
the
BMD
and
BMDL
values
changed
by
a
small
amount.
Where
appropriate,
these
revised
values
were
used
to
calculate
health
advisories
and
RfDs.

Data
for
kidney
cytomegaly
from
the
chronic
NTP
(
1987)
study
in
male
mice
were
remodeled
using
the
BMDS
program.
The
results
from
the
logistic
and
probit
models
were
rejected
for
lack
of
fit
(
p
<
0.10).
Estimates
of
the
BMDL
10
calculated
by
the
Weibull
and
multistage
models
were
identical
to
estimates
derived
using
the
Crump
benchmark
dose
software
(
0.96
mg/
kg­
day).
The
results
for
the
BMDL
10
varied
by
more
than
a
factor
of
3,
indicating
a
A
­
38
Draft
­
Do
not
cite
or
quote
February
20,
2002
degree
of
model
dependence.
The
log­
logistic
model
gave
a
very
low
value
for
the
BMDL
10
and
was
eliminated
from
further
consideration.
Of
the
remaining
models,
the
log­
probit
model
gave
the
lowest
value
for
the
AIC
and
the
corresponding
BMDL
10
was
thus
selected
as
a
candidate
for
derivation
of
the
RfD.

When
data
for
fatty
degeneration
in
the
liver
of
female
rats
(
Aida
et
al.,
1992b)
were
remodeled
using
the
BMDS
program,
all
models
fit
the
data
adequately.
The
resulting
BMDL
10
values
were
within
a
factor
of
three,
indicating
model
independence.
The
BMDL
10
calculated
using
the
probit
model
was
selected
as
a
candidate
for
derivation
of
the
RfD
because
it
had
the
lowest
AIC
value.

f.
Modeling
of
Continuous
Endpoints
Four
continuous
data
sets
for
maternal
body
weight
gain
were
modeled
in
support
of
Health
Advisory
and
RfD
derivation:
Narotsky
et
al.
(
1997);
CCC
(
2000b)
and
CCC
(
2000d).
Data
fitting
problems
were
encountered
when
attempting
to
model
these
data
sets
using
BMDS
Version
1.2.
Therefore,
these
data
sets
were
modeled
using
BMDS
Version
1.3.
This
version
was
not
available
when
the
analysis
of
dichotomous
data
sets
for
other
endpoints
was
performed.

For
the
Narotsky
et
al.
(
1997)
study,
body
weight
gain
data
for
gestation
days
6
to
8
were
modeled.
To
facilitate
modeling,
a
constant
value
of
20
was
added
to
each
mean
so
that
all
modeled
data
were
positive.
This
procedure
is
considered
an
acceptable
approach
for
transforming
data
prior
to
modeling
continuous
data
with
the
BMDS
software
(
W.
Setzer,
U.
S.
EPA,
personal
communication).
The
BMDS
tests
for
variance
rejected
the
hypothesis
that
there
is
a
constant
variance
for
this
data
set.
The
modeled
variance
available
in
BMDS
did
a
good
job
of
describing
the
variation
in
the
variances
(
p­
value
of
0.76),
when
a
constant
value
of
20
was
added
to
the
means.

When
the
models
were
fit
to
the
transformed
data,
modeling
the
nonconstant
variance
in
terms
of
the
means
plus
a
constant
value
of
20,
the
fits
to
the
means
(
plus
20)
were
all
good.
Note
that
the
number
of
parameters
for
the
power
and
polynomial
models
are
misspecified
in
BMDS
(
because
the
power
hits
the
bound
of
1
and
the
polynomial
is
linear),
and
so
the
AIC
and
p­
value
for
fit
are
incorrect.
The
correct
values
can
be
found
in
the
output
for
the
linear
model.
The
results
for
the
Narotsky
et
al.
(
1997)
body
weight
gain
data
are
summarized
below:

Model
Log­
likelihood
AIC
BMD
BMDL
Power
(
linear)
­
84.84
177.68
12.0
9.0
Polynomial
­
84.84
177.68
12.0
9.0
Hill
­
82.83
177.66
18.3
10.2
Even
though
the
Hill
model
has
two
more
parameters
than
the
power
(
linear)
model,
the
decrease
in
the
log­
likelihood
gained
by
those
extra
parameters
is
enough
to
give
the
slight
edge
A
­
39
Draft
­
Do
not
cite
or
quote
February
20,
2002
to
the
Hill
model
in
terms
of
AIC.
Thus,
the
model
of
choice
is
the
Hill
model,
since
the
BMDs
were
within
a
factor
of
3
of
one
another).

To
validate
the
choice
of
the
constant
used
in
the
modeling
of
this
data
set,
we
investigated
the
effect
of
adding
different
constants
to
the
means,
with
respect
to
the
estimates
from
the
Hill
model.
Different
constants
added
to
the
mean
change
the
parameter
estimates
obtained
in
the
maximization
of
the
likelihood.
In
fact,
the
choice
of
the
constant
can
be
viewed
as
the
determination
of
another
parameter
that
gives
the
best
fit
of
the
model
to
the
data
 
in
this
case
allowing
the
model
for
the
variance
to
be
improved.
Note
that
benchmark
responses
(
BMRs)
defined
in
terms
of
a
change
in
the
mean
equal
to
some
multiple
of
the
control
standard
deviation
will
be
appropriate
even
with
the
transformed
data,
because
adding
a
constant
to
a
set
of
observations
does
not
alter
the
standard
deviation
of
the
transformed
data.
Thus,
the
choice
of
the
BMR
defined
as
1.1
standard
deviations
is
consistent
for
any
choice
of
added
constant.
(
Any
differences
in
BMDs
and
BMDLs
noted
with
different
added
constants
is
due
to
differences
in
the
fitted
model
parameters,
not
the
definition
of
the
BMR).

In
addition
to
the
added
constant
of
20
that
was
used
for
the
comparisons
above,
we
also
examined
adding
constants
of
10,
15,
25,
or
30.
The
results
of
adding
the
different
constants
on
the
outcome
of
the
Hill
model
are
summarized
below:

Constant
Added
Log­
likelihood
BMD
BMDL
10
­
82.83
18.7
11.1
15
­
82.73
18.4
10.6
20
­
82.71
18.3
10.2
25
­
83.37
19.2
10.4
30
­
84.22
19.9
8.5
Because
the
log­
likelihood
measures
the
goodness
of
fit,
it
can
be
seen
that
the
constant
of
20
is
the
best
choice
from
among
those
that
we
tried.
Since
the
changes
in
the
BMDs
and
BMDLs
are
minor
in
the
region
of
20,
we
did
not
attempt
to
fine­
tune
the
choice
of
the
constant
any
further.
For
the
Narotsky
et
al.
(
1997)
data
set,
the
Hill
model
applied
to
the
data
(
with
a
constant
of
20
added
to
the
means)
was
selected
as
the
best
basis
for
BMD
estimation.
We
confirmed
that
the
polynomial
and
power
models
(
which
both
still
defaulted
to
a
linear
form)
did
not
yield
a
log­
likelihood
as
large
as
that
from
the
Hill
model,
for
the
choice
of
20
as
the
added
constant.
The
Hill
model
yielded
a
BMD
of
18,
with
a
BMDL
of
10.

For
the
CCC
(
2000b)
study,
data
for
body
weight
gain
in
rabbits
on
gestation
days
6
to
29
(
corrected
for
gravid
uterine
weight)
were
modeled.
For
this
data
set,
the
hypothesis
of
constant
variance
could
not
be
rejected
at
the
0.05
level
(
p
=
0.20).
Thus,
for
all
of
the
modeling
considered,
we
have
assumed
constant
variance.
A
­
40
Draft
­
Do
not
cite
or
quote
February
20,
2002
The
best­
fitting
polynomial
model
was
linear.
Unfortunately,
the
linear
model
did
not
describe
the
data
well
(
p­
value
for
goodness
of
fit
less
than
0.001).
In
contrast,
both
the
power
model
and
the
Hill
model
gave
adequate
fits
to
the
data
(
p­
values
of
0.28
and
0.52,
respectively).
The
following
table
summarizes
the
outputs
for
the
various
models:

Model
Log­
likelihood
AIC
BMD
BMDL
Polynomial
(
linear)
141.10
­
276.2
35.4
29.3
Power
146.47
­
284.94
50.3
Failed
Hill
147.04
­
284.08
53.7
Failed
Even
though
the
Hill
model
provided
a
slightly
larger
log­
likelihood,
the
gain
was
not
sufficient
to
decrease
the
AIC
below
that
associated
with
the
power
model
(
the
Hill
model
uses
1
extra
parameter,
and
thus
the
comparison
of
the
AICs
says
that
the
improvement
in
the
fit
­­
the
log­
likelihood
­­
is
not
enough
to
make
up
for
the
fact
that
there
is
that
one
extra
parameter).
[
Note:
the
AICs
in
the
BMDS
output
files
are
incorrectly
calculated
because
the
log­
likelihoods
are
positive
numbers
rather
than
negative
numbers.]
The
power
model
would
be
the
model
of
choice,
given
that
the
BMDs
for
the
two
models
that
fit
the
data
are
within
a
factor
of
3.
However,
BMDS
did
not
complete
the
calculation
of
the
BMDL
for
either
the
power
or
the
Hill
model.
Nevertheless,
it
appears
that
the
magnitude
of
the
BMD
(
around
50
mg/
kg)
would
not
make
this
endpoint
the
critical
one
with
respect
to
finding
a
health­
protective
starting
point
for
RfD
determination
(
i.
e.,
compare
50
mg/
kg
to
the
BMDs
from
other
endpoints).

Two
data
sets
for
body
weight
gain
were
modeled
for
the
CCC
(
2000d)
study
in
rats.
The
decrease
in
body
weight
gain
was
most
severe
on
gestation
days
6
to
7.
Here,
as
in
the
case
of
the
Narotsky
et
al.
(
1997)
data
set,
a
constant
needed
to
be
added
to
make
the
model
of
the
variance
acceptable.
However,
this
data
set
was
problematic.
No
model
available
in
BMDS
fit
the
data,
regardless
of
how
they
were
transformed
by
adding
a
constant.
The
power
and
polynomial
models
defaulted
to
the
linear
form
which
clearly
did
not
describe
the
change
in
the
means
as
a
function
of
dose.
The
Hill
model
had
the
required
curvature
to
describe
the
pattern
of
the
means
as
a
function
of
dose,
but
the
problem
for
that
model
(
as
well
as
for
the
others)
was
that
no
good
model
of
the
variance
can
be
determined
just
by
adding
a
constant
to
observations.
The
best
constant
found
was
around
10,
which
reduced
the
log­
likelihood
for
the
fitted
model
to
­
246.17.
This
is
compared
to
the
log­
likelihood
for
the
independent
means,
independent
variances
model
of
­
240.49.
The
comparison
of
these
two
models
would
not
accept
the
fitted
model
(
pvalue
slightly
less
than
0.025).
As
can
be
seen
on
the
figure
in
the
electronic
Appendix
B
for
this
particular
choice,
the
means
are
well
fit
by
the
model,
but
the
variances
are
not
well­
modeled,
especially
in
the
control
group,
for
which
the
estimated
standard
deviation
is
greater
than
the
observed
standard
deviation.
The
BMD
for
that
run
is
17.7
and
the
BMDL
is
14.5.
These
are
the
smallest
from
among
the
runs
that
we
made
(
using
added
constants
of
between
0
and
100,
and
also
the
constant
variance
model),
for
which
the
BMDs
ranged
between
18.3
and
19,
and
the
BMDLs
ranged
between
15.3
and
15.7.
It
is
clear
that
the
BMD
and
BMDL
estimates
are
not
A
­
41
Draft
­
Do
not
cite
or
quote
February
20,
2002
especially
sensitive
to
the
choice
of
the
added
constant.
Considering
that
the
models
overpredict
the
observed
control
standard
deviation,
these
BMD
estimates
may
be
viewed
as
slight
overestimates,
if
anything.
But
because
the
fits
are
not
particularly
good,
some
caution
might
be
warranted
if
one
is
considering
using
these
results
as
the
basis
for
regulation.

To
obtain
a
more
reliable
estimate
of
the
BMD
for
decreased
body
weight
observed
in
CCC
(
2000d),
body
weight
gain
data
were
also
modeled
for
gestation
days
6
to
9.
These
data
also
required
transformation
with
a
constant.
This
was
accomplished
by
starting
the
search
for
the
constant
with
a
minimum
of
30
and
a
maximum
of
500.
In
this
case,
it
was
determined
that
a
value
of
about
250,
added
to
the
means,
produced
an
acceptable
model
of
the
change
in
variance
as
a
function
of
the
mean
(
p­
value
of
about
0.21),
and
yielded
the
largest
maximized
likelihood
for
the
power
model.
This
result
was
first
compared
to
other
choices
of
constants
around
250
(
e.
g.,
240
and
260),
then
to
values
that
were
progressively
further
away
(
e.
g.,
values
of
200
and
300,
90
and
400,
etc.).
We
did
not
fine­
tune
the
estimate
of
the
constant,
since
the
BMD
estimates
were
stable
when
the
added
constant
was
240,
250,
or
260.

The
power
model
was
selected
for
the
above
search
for
the
constant
because
it
provided
a
much
better
fit
to
the
transformed
data
than
did
the
Hill
model
or
the
polynomial
model
for
preliminary
choices
for
the
added
constant
(
30
and
55).
Because
the
final
choice
of
the
constant
was
so
different
from
the
preliminary
choices,
we
compared
the
fits
of
the
models
to
the
transformed
data
using
250
as
the
added
constant
and
the
results
are
summarized
here:

Model
Log­
likelihood
AIC
BMD
BMDL
Power
­
304.942
619.885
22.9
18.4
Polynomial
­
336.201
682.403
34.0
Failed
Hill
­
413.929
839.858
Failed
Failed
The
Hill
model
did
not
fit
the
data
at
all
(
perhaps
due
to
a
vanishingly
small
estimate
of
the
parameter
k,
which
may
be
due
to
the
fact
that
the
model
did
not
correctly
maximize
the
likelihood).
The
polynomial
model
did
not
fit
the
data
well
and
is
much
less
satisfactory
than
the
power
model
(
compare
the
AICs
in
the
table
above).
Even
for
the
power
model,
the
p­
value
for
goodness
of
fit
was
0.036,
which
is
less
than
the
standard
critical
p­
value
of
0.05.
However,
for
this
data
set
which
is
nonmonotonic,
the
power
model
does
a
satisfactory
job.
Consequently,
the
reasonable
choice
for
the
BMD
is
23
and
for
the
BMDL
is
18,
for
this
data
set.

Table
A­
5
Benchmark
Dose
Modeling
Results
for
Bromodichloromethane
(
Dichotomous
Endpoints)
A
­
42
Draft
­
Do
not
cite
or
quote
February
20,
2002
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
DEVELOPMENTAL
OR
REPRODUCTIVE
STUDIES
Bielmeier
et
al.
(
2001)
Rat
Female
Full
Litter
Resorption
Gamma
1.0
24.9923
3
(
2)
36
2.1
42
4.3
Logistic
0.77
25.0528
2
(
2)
31
8.6
40
16
Log­
logistic
1.0
24.9223
3
(
2)
40
0.8
46
1.6
Log­
probit
1.0
24.9223
3
(
2)
41
5.3
45
7.7
Multistage
0.91
23.1120
1
16
2.1
23
4.2*

Probit
0.83
24.9874
2
(
2)
28
7.9
37
15
Weibull
1.0
24.9223
3
(
2)
26
2.1
34
4.3
Narotsky
et
al.
(
1997)
Rat
Female
Full
Litter
Resorption
Gamma
0.68
30.3049
3
(
2)
36
11
49
22
Logistic
0.47
31.0964
2
(
2)
41
25
54
39
Log­
logistic
0.68
30.3427
3
(
2)
36
0.13
(?)
48
5.9
(?)

Log­
probit
0.72
30.1687
3
(
2)
36
21
48
30*

Multistage
0.68
30.4257
2
33
10
47
21
Probit
0.53
30.8108
2
(
2)
40
23
52
37
Weibull
0.67
30.3983
3
(
2)
35
10
49
22
Ruddick
et
al.
(
1983)
Rat
Female
Sternebral
Aberrations
Gamma
0.44
64.2926
3
(
3)
16
7.1
30
15
Logistic
0.66
62.5139
2
(
2)
22
14
43
28
Log­
logistic
0.47
64.2131
3
(
3)
19
Failed
33
Failed
Log­
probit
0.49
64.1729
3
(
3)
23
Failed
37
Failed
Multistage
0.74
62.2980
2
13
7.1
27
15*

Probit
0.67
62.5023
2
(
2)
22
14
42
27
Weibull
0.44
64.2965
3
(
3)
14
7.1
29
15
CANDIDATE
STUDIES
FOR
1­
DAY
HA
Lilly
et
al.
(
1994)
Rat
Male
Hepatic
midzonal
vacuolar
degeneration
(
Aqueous
vehicle)

Gamma
0.63
3.8029
3
(
1)
187
135
206
156
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
43
Draft
­
Do
not
cite
or
quote
February
20,
2002
Logistic
1.0
4.0000
2
(
2)
290
145
292
174
Log­
logistic
0.99
2.0468
3
(
1)
240
164
250
182
Log­
probit
1.0
2.0000
3
(
1)
258
168
263
182*

Multistage
1.0
(?)
298.31
2
6.4
0.1
9.2
0.28
Probit
1.0
4.0000
2
(
2)
281
184
(
W)
285
187
(
W)

Weibull
1.0
2.0006
3
(
1)
294
144
307
170
Lilly
et
al.
(
1994)
Rat
Male
Renal
Tubule
degeneration
(
Aqueous
vehicle)

Gamma
1.0
9.6389
3
(
1)
119
4.3
131
8.9*

Logistic
1.0
11.6382
2
(
2)
163
19
171
35
Log­
logistic
1.0
9.6382
3
(
1)
163
0.00078
(?)
170
0.00625
(?)

Log­
probit
1.0
11.6382
3
(
2)
152
11
160
16
Multistage
1.0
(?)
102.1000
2
6.4
0.13
9.2
0.12
Probit
1.0
11.6302
2
(
2)
132
17
144
33
Weibull
1.0
11.6302
3
(
2)
92
4.4
110
8.9
Thornton­
Manning
et
al.
(
1994)
Rat
Female
Hepatic
centrilobular
vacuolar
degeneration
(
Poor
fit:
no
model
selected
Gamma
0.01
19.0705
3
(
2)
40
5.5
53
11
Logistic
<
0.01
21.4063
2
(
2)
34
16
54
29
Log­
logistic
<
0.01
17.3420
3
(
2)
55
13
67
22
Log­
probit
0.02
17.9809
3
(
2)
53
16
64
22
Multistage
0.01
20.1038
2
17
5.0
31
10
Probit
<
0.01
219.846
2
(
2)
31
15
52
29
Weibull
0.02
19.6598
3
(
2)
23
5.1
37
10
Thornton­
Manning
et
al.
(
1994)
Rat
Female
Renal
tubular
degeneration
Gamma
1.0
10.3742
3
(
1)
99
45
109
60
Logistic
1
12.3178
2
(
2)
138
42
141
63
Log­
logistic
1
10.3178
3
(
1)
127
51
133
65*

Log­
probit
1
12.3178
3
(
2)
123
54
128
66
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
44
Draft
­
Do
not
cite
or
quote
February
20,
2002
Multistage
1
(?)
445.4650
2
3.4
0.06
4.4
0.08
Probit
1
12.3178
2
(
2)
128
38
133
58
Weibull
1
12.3179
3
(
2)
127
39
133
56
CANDIDATE
STUDIES
FOR
THE
10­
DAY
HA
Condie
et
al.
(
1983)
Mouse
Male
Hepatic
centrilobular
pallor
Gamma
0.19
30.0268
3
(
2)
11
2.2
17
4.5
Logistic
<
0.01
33.9633
2
(
2)
14
7.9
23
14
Log­
logistic
0.30
28.7294
3
(
2)
19
4.4
24
7.5*

Log­
probit
0.26
29.2521
3
(
2)
18
6.0
23
8.6
Multistage
0.16
30.6962
2
4.2
2.1
8.5
4.3
Probit
<
0.01
34.7900
2
(
2)
13
7.6
22
14
Weibull
0.18
30.3429
3
(
2)
7.2
2.1
12
4.4
Condie
et
al.
(
1983)
Mouse
Male
Renal
epithelial
hyperplasia
Gamma
0.88
35.6018
3
(
2)
75
46
83
56
Logistic
0.10
40.0819
2
(
2)
25
15
40
27
Log­
logistic
0.98
35.3785
3
(
2)
112
48
117
58
Log­
probit
0.84
37.3705
3
(
3)
109
51
113
59
Multistage
0.38
37.5968
2
46
17
58
31
Probit
0.08
40.7541
2
(
2)
21
13
35
24
Weibull
0.98
35.3785
3
(
2)
120
41
125
53*

Aida
et
al.
(
1992a)
Rat
Female
Liver
cell
vacuolization
Gamma
0.56
23.4101
3
(
2)
18
8.5
36
17
Logistic
0.23
25.2544
2
(
2)
50
27
83
49
Log­
logistic
0.60
23.2879
3
(
2)
19
0.0015
(?)
35
0.1
(?)

Log­
probit
0.65
21.5114
3
(
2)
34
20
49
28
Multistage
0.78
21.4146
1
16
8.5
34
17*

Probit
0.25
25.0481
2
(
2)
46
25
76
46
Weibull
0.57
23.4143
3
(
2)
17
8.5
34
17
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
45
Draft
­
Do
not
cite
or
quote
February
20,
2002
Melnick
et
al.
(
1998)
Mouse
Female
Hepatocyte
hydropic
degeneration
Gamma
0.99
26.2625
3
(
2)
28
4.2
35
8.5
Logistic
0.82
26.8455
2
(
2)
24
11
36
20
Log­
logistic
0.91
26.5313
3
(
2)
31
11
38
16
Log­
probit
0.96
26.3532
3
(
2)
32
12
38
17
Multistage
1.0?
445.4650
2
3.4
0.076
4.4
0.082
Probit
0.89
26.5905
3(
2)
24
11
34
19
Weibull
1.0
26.2278
3(
2)
23
4.2
31
8.4*

NTP
(
1998)
Rat
Male
Single
cell
hepatic
necrosis
Gamma
1.0
14.4941
3
(
1)
26
12
28
15.125
Logistic
1.0
16.3653
2
(
2)
34
12
35
16.9848
Log­
logistic
1.0
14.3662
3
(
1)
33
15
34
18.4508*

Log­
probit
1.0
16.3653
3
(
2)
33
15
334
17.4679
Multistage
0.93
15.1127
1
16
4.8
20
9.4
Probit
1.000
16.3653
2
(
2)
31
10
32
15
Weibull
1.000
16.3653
3
(
2)
30
9.8
32
14
CANDIDATE
STUDIES
FOR
THE
LONGER­
TERM
HA
NTP
(
1987)
Mouse
Female
Hepatic
Vacuolated
Cytoplasm
Gamma
0.27
32.2103
3
(
2)
57
28
73
42
Logistic
0.04
37.1630
2
(
2)
54
33
80
54
Log­
logistic
0.38
31.1109
3
(
2)
60
33
74
46
Log­
probit
0.41
30.9082
3
(
2)
62
35
75
47*

Multistage
0.32
31.6422
1
44
20
63
37
Probit
0.05
36.4925
2
(
2)
55
32
80
54
Weibull
0.20
33.6311
3
(
2)
46
21
65
35
CANDIDATE
STUDIES
FOR
THE
RfD
Aida
et
al.
(
1992b)
Rat
Female
Hepatic
Fatty
Degeneration
Gamma
1.0
46.8266
­­
5.1
0.71
5.8
1.4
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
46
Draft
­
Do
not
cite
or
quote
February
20,
2002
Logistic
0.99
44.8652
­­
2.0
1.3
3.4
2.3
Log­
logistic
1.0
42.8266
­­
6.7
2.0
7.1
2.9
Log­
probit
1.0
46.8266
­­
5.9
2.0
6.4
2.7
Multistage
1.0
48.8266
­­
2.9
0.56
4.0
1.1
Probit
1.0
44.8296
­­
1.8
1.2
3.1
2.1*

Weibull
1.0
46.8266
­­
3.3
0.65
4.4
1.3
Aida
et
al.
(
1992b)
Rat
Male
Hepatic
Fatty
Degeneration
Gamma
1.0
29.3001
3
(
2)
1.2
0.39
2.1
0.80
Logistic
0.15
34.2633
2
(
2)
2.6
1.5
4.4
2.7
Log­
logistic
0.98
29.3760
3
(
2)
2.1
0.51
3.0
0.94
Log­
probit
1.0
29.3074
3
(
2)
2.2
0.99
3.0
1.4
Multistage
1.0
29.3001
2
0.73
0.43
1.5
0.88
Probit
0.17
33.8989
2
(
2)
2.6
1.6
4.4
2.9
Weibull
1.0
29.3001
3
(
2)
1.1
0.39
1.9
0.80*

Aida
et
al.
(
1992b)
Rat
Male
Hepatic
Granulomas
Gamma
0.99
34.9241
3
(
1)
1.0
0.67
2.1
1.4
Logistic
0.13
42.0098
2
(
2)
4.2
2.6
7.0
4.6
Log­
logistic
0.66
38.0328
3
(
2)
1.9
0.37
3.0
0.82
Log­
probit
0.87
35.6192
3
(
1)
2.4
1.6
3.5
2.3
Multistage
0.95
36.8960
2
1.1
0.67
2.2
1.4
Probit
0.15
41.5172
2
(
2)
3.8
2.4
6.4
4.4
Weibull
0.99
34.9241
3
(
1)
1.0
0.67
2.1
1.4*

NTP
(
1987)
Mouse
Male
Hepatic
Focal
Necrosis
Gamma
1.0
15.6352
3
(
1)
44
27
49
34
Logistic
1.0
17.4602
2
(
2)
65
30
66
38
Log­
logistic
1.0
15.4604
3
(
1)
59
28
61
34
Log­
probit
1.0
17.4602
3
(
2)
57
28
60
34
Multistage
1.0
16.0769
1
40
23
47
32
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
47
Draft
­
Do
not
cite
or
quote
February
20,
2002
Probit
1.0
17.4602
2
(
2)
60
28
62
36
Weibull
1.0
15.4603
3
(
1)
60
27
63
35*

NTP
(
1987)
Mouse
Male
Kidney
Cytomegaly
Gamma
0.76
72.3705
0.58
0.47
1.2
0.96
Logistic
<
0.01
88.5494
3.3
2.1
5.3
3.8
Log­
logistic
1.0
73.8361
1.5
3.6E­
9
2.2
1.1E­
7
Log­
probit
0.99
71.8572
1.4
1.1
2.0
1.5*

Multistage
0.45
74.3705
0.58
0.47
1.2
0.96
Probit
<
0.01
92.6760
2.8
2.0
4.8
3.6
Weibull
0.76
72.3705
0.58
0.47
1.2
0.96
*
Selected
model
result
for
endpoint.
AIC
Akaike
Information
Criterion
BMD
Benchmark
Dose
BMDL
95%
lower
confidence
level
on
BMD
G­
O­
F
Goodness­
of­
Fit
HA
Health
Advisory
?
Results
questionable
on
the
basis
of
visual
inspection
or
probable
calculation
error
W
BMDS
gave
a
warning
message:
"
BMDL
computation
is
at
best
imprecise
for
these
data"
A
­
48
Draft
­
Do
not
cite
or
quote
February
20,
2002
2.
Dibromochloromethane
BMDS
modeling
results
for
dibromochloromethane
are
summarized
in
Table
A­
6
below.
Detailed
output
for
each
model
run
is
compiled
in
Appendix
B,
provided
in
electronic
format
on
compact
disk.

a.
Developmental
and
Reproductive
Studies
No
data
were
modeled.
The
generation
F2b
day
14
postnatal
body
weight
data
of
Borzelleca
and
Carchman
(
1982)
were
considered
for
modeling.
However,
the
study
authors
did
not
report
the
number
of
litters
examined
for
this
continuous
endpoint.
Since
this
information
is
required
as
input,
the
data
could
not
be
modeled.

b.
One­
day
Health
Advisory
No
data
were
modeled
for
the
One­
day
HA.

c.
Ten­
day
Health
Advisory
Seven
data
sets
were
modeled
in
support
of
the
Ten­
day
HA.
When
data
were
analyzed
for
the
liver
cell
vacuolation
in
female
rats
(
Aida1992a),
the
multistage
model
gave
questionable
results
and
was
eliminated
from
consideration.
The
remaining
BMDL
values
were
within
a
factor
of
3,
so
the
estimate
from
the
Weibull
and
gamma
models
was
selected
on
the
basis
of
the
smallest
AIC.
For
the
same
endpoint
in
male
rats
(
Aida
et
al.
1992a),
model
fit
was
adequate
in
all
cases
and
all
BMDL
values
were
within
a
factor
of
3.
The
multistage
model
was
selected
on
the
basis
of
the
smallest
AIC.

Liver
and
kidney
histopathology
data
from
the
study
by
Condie
et
al.
(
1983)
were
analyzed.
In
the
analysis
of
kidney
data,
the
BMDLs
calculated
by
the
logistic
and
probit
models
were
eliminated
because
the
models
fit
the
data
poorly.
Among
the
remaining
models,
the
loglogistic
BMDLs
are
smallest
by
more
than
a
factor
of
3.
Thus,
this
result
was
examined
as
a
possible
outlier.
Although
this
model
does
at
times
have
difficulty
calculating
reasonable
lower
bounds,
the
BMDs
in
this
case
are
also
smaller
than
the
BMDs
from
the
other
models
(
although
not
by
a
factor
of
3
for
all
them
 
the
BMDLs
calculated
by
the
logistic
model
are
more
than
3­
fold
below
any
other
BMDL).
In
addition,
the
AIC
value
is
much
lower
for
log­
logistic
than
for
the
other
models.
Thus,
the
BMDL
calculated
by
the
log­
logistic
model
was
selected.

With
respect
to
the
Condie
et
al.
(
1983)
liver
histopathology
data,
the
log­
logistic
model
gives
the
smallest
BMDL
by
more
than
a
factor
of
3.
However,
the
BMDs
are
within
a
factor
of
3
of
the
BMDs
generated
by
other
models
and
the
AIC
for
the
log­
logistic
model
is
not
the
lowest.
The
log­
logistic
results
were
therefore
considered
as
outliers.
Among
the
remaining
options,
the
Weibull
and
gamma
models
have
the
lowest
AIC,
and
the
BMDL
calculated
by
these
models
was
selected.
A
­
49
Draft
­
Do
not
cite
or
quote
February
20,
2002
Data
for
stomach
nodules
in
male
and
female
rats
(
NTP,
1985)
were
also
analyzed.
In
females,
all
calculated
BMDLs
are
within
a
factor
of
3,
so
the
result
from
the
multistage
model
was
selected
on
the
basis
of
having
the
smallest
AIC.
In
males,
the
log­
logistic
model
has
the
smallest
AIC
and
the
lowest
BMDL.

In
the
analysis
of
the
Melnick
et
al.
(
1998)
results
for
hepatic
hydropic
degeneration
in
female
mice,
the
multistage
model
failed
to
fit
the
data.
All
remaining
results
were
similar
and
the
BMDL
calculated
by
the
log­
logistic
model
was
selected
on
the
basis
of
the
lowest
AIC.

d.
Longer­
term
Health
Advisory
Two
data
sets
for
hepatic
lesions
reported
in
Chu
et
al.
(
1982b)
were
modeled
in
support
of
the
Longer­
term
HA.
When
liver
histopathology
data
for
male
rats
was
analyzed,
the
loglogistic
model
calculated
a
BMDL
that
was
more
than
3­
fold
lower
than
those
from
some
other
models.
However,
both
the
AIC
and
the
BMD
calculated
by
the
log­
logistic
model
were
the
lowest
among
all
models.
The
BMDL
calculated
by
this
model
was
thus
selected.
When
data
for
liver
lesions
in
female
rats
were
analyzed,
no
model
adequately
fit
the
data
(
all
p
values
were
less
than
0.1).
Thus,
no
BMDL
was
selected
from
this
data
set.

Data
for
fatty
metamorphosis
in
the
liver
of
male
rats
that
had
previously
been
modeled
using
the
Crump
software
was
reanalyzed
using
the
BMDS
program.
All
models
adequately
fit
the
data
for
this
endpoint.
All
resulting
BMDL
values
were
within
a
factor
of
three,
with
the
exception
of
the
estimate
calculated
using
the
log­
probit
model
The
value
calculated
using
the
probit
model
was
selected
on
the
basis
of
the
lowest
AIC.

e.
RfD
Two
data
sets
from
the
NTP
(
1985)
chronic
oral
exposure
study
were
modeled
using
the
BMDS
software
in
support
of
the
RfD
for
dibromochloromethane.
These
data
sets
were
selected
after
inspection
of
the
results
for
BMD
modeling
of
key
dibromochloromethane
endpoints
using
the
Crump
software
(
K.
S.
Crump,
Inc.).
For
fatty
metamorphosis
in
the
liver
of
male
rats,
all
models
fit
the
data
acceptably,
although
the
p
value
for
the
probit
model
was
marginal
(
p
=
0.16).
All
BMDL
values
were
within
a
factor
of
three,
with
the
exception
of
the
log­
logistic
model.
Results
from
the
log­
logistic
model
were
selected
on
the
basis
of
the
lowest
AIC
value.
For
ground
glass
cytoplasm
in
the
liver
of
male
rats,
all
models
fit
the
data
acceptably
and
the
BMDL
values
were
within
a
factor
of
three.
The
results
for
the
probit
model
were
selected
on
the
basis
of
the
lowest
AIC.
A
­
50
Draft
­
Do
not
cite
or
quote
February
20,
2002
Table
A­
6
Benchmark
Dose
Modeling
Results
for
Dibromochloromethane
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
DEVELOPMENTAL
AND
REPRODUCTIVE
STUDIES
(
NONE)

CANDIDATE
STUDIES
FOR
1­
DAY
HA
(
NONE)

CANDIDATE
STUDIES
FOR
10­
DAY
HA
Aida
et
al.
(
1992a)
Rat
Female
Liver
cell
vacuolization
Gamma
1.0
15.4833
3
(
2)
24
3.2
30
6.7
Logistic
0.91
15.7862
2
(
2)
24
10
34
17
Log­
logistic
0.99
15.5265
3
(
2)
24
7.5
30
12
Log­
probit
1.0
15.4078
3
(
2)
25
8.4
30
12
Multistage
1
(?)
347
2
3.4
0.063
4.4
0.064
Probit
0.95
15.6586
2
(
2)
22
8.9
32
16
Weibull
1.0
15.4833
3
(
2)
21
3.2
29
6.7*

Aida
et
al.
(
1992a)
Rat
Males
Liver
cell
vacuolization
Gamma
0.76
20.0153
3
(
2)
12
2.5
18
5.1
Logistic
0.77
20.0459
2
(
2)
17
8.0
26
14
Log­
logistic
0.56
20.8287
3
(
2)
14
2.7
20
5.3
Log­
probit
0.57
20.7294
3
(
2)
14
6.0
19
8.6
Multistage
0.98
19.3422
2
7.0
2.7
14
5.5*

Probit
0.80
19.9157
2
(
2)
15
7.4
24
13
Weibull
0.83
19.7280
3
(
2)
12
2.6
18
5.3
Condie
et
al
(
1983)
Mouse
Male
Renal
mesangial
hypertrophy
Gamma
0.12
36.8675
3
(
1)
3.8
2.6
7.8
5.3
Logistic
<
0.01
47.0265
2
(
2)
12
7.7
22
15
Log­
logistic
0.49
33.8237
3
(
1)
16
0.7
3.5
1.6*

Log­
probit
0.11
36.7799
3
(
1)
8.9
5.6
13
8.1
Multistage
0.12
36.8675
2
(
1)
3.8
2.6
7.8
5.3
Probit
<
0.01
46.9299
2
(
2)
12
8.0
22
16
Weibull
0.12
36.8675
3
(
1)
3.8
2.6
7.8
5.3
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
51
Draft
­
Do
not
cite
or
quote
February
20,
2002
Condie
et
al.
(
1983)
Mouse
Male
hepatic
cytoplasmic
vacuolization
Gamma
0.32
43.8699
3
(
2)
5.4
3.4
11
6.9
Logistic
0.15
45.1424
2
(
2)
15
9.4
27
18
Log­
logistic
0.33
43.9124
3
(
2)
3.3
1.5
7.0
3.3
Log­
probit
0.23
44.4103
3
(
2)
14
8.5
20
12
Multistage
0.13
45.8578
3
5.7
3.4
12
6.9
Probit
0.16
44.9950
2
(
2)
14
9.2
26
18
Weibull
0.32
43.8699
3
(
2)
5.4
3.4
11
6.9*

NTP
(
1985)
Mouse
Female
Stomach
nodules
Gamma
0.98
16.2743
3
(
2)
167
38
225
78
Logistic
0.86
17.1462
2
(
2)
209
104
284
170
Log­
logistic
0.98
16.2955
3
(
2)
163
34
222
73
Log­
probit
0.99
16.1399
3
(
2)
166
74
218
106
Multistage
0.99
14.3879
1
152
37
218
77*

Probit
0.90
16.8628
2
(
2)
197
95
267
158
Weibull
0.97
16.3761
3
(
2)
162
37
227
77
NTP
(
1985)
Mouse
Male
Stomach
nodules
Gamma
0.91
18.5058
3
(
1)
75
33
153
67
Logistic
0.64
21.8294
2
(
2)
191
97
306
168
Log­
logistic
0.93
18.5021
3
(
1)
68
26
143
54*

Log­
probit
0.68
19.4319
3
(
1)
122
69
176
99
Multistage
0.91
18.5858
1
75
33
153
67
Probit
0.65
21.7110
2
(
2)
174
88
284
154
Weibull
0.91
18.5858
3
(
1)
75
33
153
67
Melnick
et
al
(
1998)
Mouse
Female
hepatic
hydropic
degeneration
Gamma
0.99
12.6487
3
(
1)
76
55
84
64
Logistic
1.0
14.0080
2
(
2)
123
56
126
70
Log­
logistic
1.0
12.0080
3
(
1)
108
59
112
68*
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
52
Draft
­
Do
not
cite
or
quote
February
20,
2002
Log­
probit
1.0
14.0080
3
(
2)
107
60
111
68
Multistage
program
failed!
­­
­­
­­
­­
­­

Probit
1.0
14.0080
2
(
2)
112
56
116
68
Weibull
1.0
14.0080
3
(
2)
112
56
117
68
CANDIDATE
STUDIES
FOR
LONGER­
TERM
HA
Chu
et
al.
(
1982b)
Rat
Male
Hepatic
lesions
Gamma
0.84
65.3956
3
(
2)
14
6.2
29
13
Logistic
0.79
65.6217
2
(
2)
22
12
43
23
Log­
logistic
0.88
65.1734
3
(
2)
8.6
2.5
18
5.3*

Log­
probit
0.72
65.8723
3
(
2)
37
15
54
22
Multistage
0.84
65.3956
2
14
6.2
29
13
Probit
0.79
65.6176
2
(
2)
22
12
43
24
Weibull
0.84
65.3956
3
(
2)
14
6.2
29
13
Chu
et
al.
(
1982b)
Rat
Female
Hepatic
lesions
Gamma
0.04
67.0864
3
(
3)
Flat
Curve
Estimated
No
BMDs
­­

Logistic
0.09
64.3869
2
(
2)
66
26
127
49
Log­
logistic
0.09
64.3474
3
(
2)
48
9.8
101
21
Log­
probit
0.04
67.0864
3
(
3)
4800
38
6900
54
Multistage
0.09
64.3615
2
54
15
110
30
Probit
0.09
64.3843
2
(
2)
64
25
125
48
Weibull
0.09
64.3615
3
(
2)
54
15
110
30
NTP
1985
Rat
Male
Fatty
Metamorphosis
(
Subchronic)

Gamma
0.90
42.3900
3
(
3)
2.6
0.44
3.9
0.91
Logistic
0.97
40.3442
2
(
2)
1.2
0.76
2.4
1.5
Log­
logistic
0.81
42.9172
3
(
3)
4.1
0.20
5.5
0.42
Log­
probit
0.85
42.6546
3
(
3)
4.3
1.1
5.4
1.6
Multistage
0.92
43.8670
4
1.0
0.49
2.1
1.0
Probit
0.98
40.1651
2
(
2)
1.3
0.84
2.5
1.7*
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
53
Draft
­
Do
not
cite
or
quote
February
20,
2002
Weibull
0.92
42.2885
3
(
3)
2.4
0.45
3.9
0.92
CANDIDATE
STUDIES
FOR
THE
RfD
NTP
(
1985)
Rat
Male
Hepatic
Fatty
Metamorphosis
­
Chronic
Gamma
0.53
105.8690
3
(
2)
0.81
0.57
1.7
1.2
Logistic
0.34
106.2850
2
(
2)
1.2
0.86
2.4
1.7
Log­
logistic
1.0
107.4950
3
(
3)
1.7
0.071
2.7
0.15
Log­
probit
0.86
105.5270
3
(
2)
1.9
1.1
2.7
1.6*

Multistage
0.53
105.8690
2
0.81
0.57
1.7
1.2
Probit
0.16
107.2230
2
(
2)
1.4
1.1
2.8
2.2
Weibull
0.53
105.8690
3
(
2)
0.81
0.57
1.7
1.2
NTP
(
1985)
Rat
Male
Ground
Glass
Cytoplasm
­
Chronic
Gamma
1.0
181.2470
3
(
3)
6.1
2.4
10
5.0
Logistic
0.58
179.5560
2
(
2)
6.3
5.1
12
9.7
Log­
logistic
1.0
181.2470
3
(
3)
7.6
1.6
12
3.5
Log­
probit
1.0
181.2470
3
(
3)
9.1
6.3
13
9.1
Multistage
1.0
181.2470
3
4.3
2.4
8.6
5.0
Probit
0.62
179.4870
2
(
2)
6.0
4.9
11
9.4*

Weibull
1.0
181.2470
3
(
3)
5.5
2.4
9.8
5.0
*
Selected
model
result
for
endpoint.

AIC
Akaike
Information
Criterion
BMD
Benchmark
Dose
BMDL
95%
lower
confidence
level
on
BMD
G­
O­
F
Goodness­
of­
Fit
HA
Health
Advisory
?
Results
questionable
on
the
basis
of
visual
inspection
or
probable
calculation
error
W
BMDS
gave
a
warning
message:
"
BMDL
computation
is
at
best
imprecise
for
these
data"
A
­
54
Draft
­
Do
not
cite
or
quote
February
20,
2002
3.
Bromoform
BMDS
modeling
results
for
dibromochloromethane
are
summarized
in
Table
A­
7
below.
Detailed
output
for
each
model
run
is
compiled
in
Appendix
B,
provided
in
electronic
format
on
compact
disk.

a.
Developmental
and
Reproductive
Studies
Data
from
the
study
by
Ruddick
et
al.
(
1983)
consisted
of
the
count
of
the
numbers
of
litters
that
had
one
or
more
fetuses
with
sternebral
variations.
This
expression
of
the
response
rates
does
not
correspond
directly
to
the
probability
of
a
response
in
the
offspring
of
treated
dams.
All
of
the
model
fit
the
data
and
all
of
the
BMDL
results
for
10%
extra
risk
are
within
a
factor
of
three
of
one
another,
so
the
results
from
the
log­
probit
model
were
selected,
because
that
model
had
the
lowest
AIC.

b.
One­
day
Health
Advisory
BMD
calculations
were
not
conducted
in
support
of
the
One­
day
HA
due
to
a
lack
of
appropriate
data.

c.
Ten­
day
Health
Advisory
Six
data
sets
were
modeled
in
support
of
the
Ten­
day
HA.
Modeling
results
for
each
data
set
were
evaluated
using
the
criteria
given
in
Section
C.
For
the
Aida
et
al.
(
1992a)
data
on
liver
cell
vacuolation
in
female
rats,
all
models
fit
well
and
(
with
one
exception)
give
the
same
AIC.
The
log­
logistic
results
appear
to
be
qualitative
outliers
because
they
are
more
than
3
times
less
than
the
next
closest
BMDLs,
even
though
the
BMD
is
the
second
largest.
The
BMDL
calculated
by
this
model
was
thus
rejected
in
favor
of
the
next
lowest
BMDL
(
Weibull
model).
When
the
same
endpoint
was
modeled
in
male
rats
(
Aida
et
al.
1992),
the
probit
model
either
failed
or
gave
a
warning
message
for
the
lower
bound
calculations
and
results
were
thus
eliminated.
The
remaining
models
calculated
very
similar
BMDLs.
The
Weibull
model
was
selected
because
it
gave
the
lowest
AIC
among
the
remaining
models.

For
histopathological
effects
in
the
kidney
of
male
mice
(
Condie
et
al.,
1983),
two
models
(
logistic
and
probit)
had
somewhat
higher
BMDLs
and
the
largest
values
for
AIC.
If
these
results
are
eliminated
as
qualitative
outliers,
the
remaining
BMDLs
are
within
a
factor
of
3.
The
Log­
probit
model
was
selected
from
among
the
remaining
models
because
it
gave
the
lowest
AIC.
Modeling
of
data
on
liver
histopathology
from
the
same
study
gave
a
similar
pattern
of
results.
Results
from
the
probit
and
logistic
models
were
eliminates
as
qualitative
outliers
(
high
AICs
and
BMDLs
that
were
higher
by
more
than
a
factor
of
3
from
the
lowest
BMDL).
The
remaining
BMDLs
are
within
a
factor
of
3
and
so
the
Log­
probit
model
was
selected
because
it
had
the
lowest
AIC.

Melnick
et
al.
(
1998)
reported
data
for
hydropic
degeneration
in
the
liver
of
female
mice.
The
multistage
model
gave
questionable
results
(
very
high
AIC
and
a
goodness
of
fit
p
value
that
A
­
55
Draft
­
Do
not
cite
or
quote
February
20,
2002
appeared
unrealistically
high
when
the
model
fit
was
examined
visually)
for
this
data
set
that
appeared
to
reflect
a
calculation
error
in
the
BMDS
software.
The
BMDLs
estimated
by
the
remaining
models
are
very
close.
The
Log­
probit
model
was
selected
because
it
has
the
lowest
AIC.
When
data
for
stomach
nodules
in
male
mice
were
modeled
(
NTP,
1989a),
all
models
gave
an
acceptable
fit
and
all
BMDLS
were
within
a
factor
of
three.
The
results
from
the
multistage
model
were
selected
because
it
had
the
lowest
AIC.

d.
Longer­
term
Health
Advisory
Three
data
sets
were
modeled
in
support
of
the
longer­
term
Health
Advisory
for
Bromoform.
No
models
adequately
fit
(
i.
e.
all
p
values
for
goodness
of
fit
were
less
than
0.1)
the
Chu
et
al.
(
1982b)
data
for
hepatic
lesions
in
female
rats
(
nonmonotonic
dose
response).
Thus,
none
of
the
calculated
BMDLs
were
candidates
for
deriving
the
Longer­
term
Health
Advisory.
For
the
same
endpoint
in
male
rats
(
Chu
et
al.
1982b),
the
multistage
model
gave
a
bad
fit
and
was
eliminated
from
consideration.
The
log­
logistic
BMDLs
are
the
lowest
of
the
remaining
values,
but
there
is
a
spread
of
greater
than
3.
The
result
for
the
log­
logistic
model
was
eliminated
as
a
qualitative
outlier,
since
this
model
gave
the
largest
BMDs
but
the
BMDLs
were
among
the
lowest
observed
(
i.
e.
gave
a
wide
confidence
interval).
The
remaining
BMDL
values
were
similar
and
the
probit
model
was
selected
because
it
gave
the
lowest
AIC
value.
Modeling
of
the
NTP
(
1989a)
data
for
hepatic
vacuolation
in
female
mice
gave
similar
results
across
all
models.
The
Log­
probit
model
was
selected
because
it
gave
the
lowest
AIC
.

e.
RfD
Two
data
sets
from
the
oral
exposure
study
conducted
in
rats
by
NTP
(
1989a)
were
modeled
using
the
BMDS
program
for
consideration
in
derivation
of
the
RfD.
For
hepatic
vacuolization
in
male
rats
exposed
to
bromoform
for
13
weeks,
no
fit
was
obtained
for
the
multistage
model.
The
BMDL
values
calculated
using
the
remaining
models
were
with
a
factor
of
3,
with
the
exception
of
the
log­
logistic
model.
The
results
from
the
Weibull
model
were
selected
on
the
basis
of
the
lowest
AIC
value.
With
respect
to
data
for
fatty
changes
in
the
liver
of
male
rats
chronically
exposed
to
bromoform
(
NTP,
1989a),
all
models
gave
acceptable
fits.
BMDLs
calculated
by
all
models
except
the
log­
logistic
were
within
a
factor
of
3.
The
log­
logistic
model
was
eliminated
as
a
qualitative
outlier,
since
it
gave
the
highest
BMDs
but
very
low
BMDLs
(
i.
e.
it
resulted
in
a
very
wide
confidence
interval).
Of
the
remaining
models,
the
lowest
AIC
was
observed
for
the
multistage
and
it
was
therefore
selected.
A
­
56
Draft
­
Do
not
cite
or
quote
February
20,
2002
Table
A­
7
Benchmark
Dose
Modeling
Results
for
Bromoform
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
REPRODUCTIVE
STUDIES
Ruddick
et
al.
(
1983)
Rat
Females
Sternebral
Aberrations
Gamma
0.85
67.302
3
(
3)
16
9.2
32
19
Logistic
0.70
66.0887
2
(
2)
33
23
59
42
Log­
logistic
0.90
67.3517
3
(
3)
19
6.4
35
14
Log­
probit
0.87
65.6053
3
(
2)
35
23
50
33*

Multistage
0.85
67.339
3
15
9.1
31
19
Probit
0.74
65.9551
2
(
2)
30
21
55
40
Weibull
0.85
67.3711
3
(
3)
16
9.2
32
19
CANDIDATE
STUDIES
FOR
1­
DAY
HA
(
NONE)

CANDIDATE
STUDIES
FOR
10­
DAY
HA
Aida
et
al.
(
1992a)
Rat
Females
Liver
cell
vacuolization
Gamma
1.0
12.3758
3(
2)
23
1.1
28
2.3
Logistic
1.0
12.3758
2
(
2)
45
5.2
47
9.6
Log­
logistic
1.0
12.3758
3
(
2)
42
0.29
45
0.61
Log­
probit
1.0
12.3758
3
(
2)
32
2.6
35
3.7
Multistage
1.0
14.3758
3
9.1
1.9
14
2.4
Probit
1.0
12.3758
2
(
2)
36
4.8
40
9.0
Weibull
1.0
12.3758
3
(
2)
11
1.1
16
2.3*

Aida
et
al.
(
1992a)
Rat
Males
Liver
cell
vacuolization
Gamma
0.99
2.2426
3
(
1)
73
44
81
53
Logistic
1.0
4.0000
2
(
2)
116
56
118
57
Log­
logistic
1.0
2.0014
3
(
1)
91
48
95
56
Log­
probit
1.0
4.0000
3
(
2)
95
49
93
56
Multistage
0.76
4.2093
1
46
14
59
28
Probit
1.0
4.0000
2
(
2)
118
failed
121
58
(
W)

Weibull
1.0
2.0000
3
(
1)
134
40
140
51*
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
57
Draft
­
Do
not
cite
or
quote
February
20,
2002
Condie
et
al.
(
1983)
Mouse
Male
Renal
mesangial
nephrosis
Gamma
0.46
32.2294
3
(
2)
45
9.1
64
19
Logistic
0.13
38.0766
2
(
2)
54
30
85
54
Log­
logistic
0.51
31.9498
3
(
2)
49
6.6
68
14
Log­
probit
0.53
31.8168
3
(
2)
57
23
73
34*

Multistage
0.40
32.6120
2
29
8.7
51
18
Probit
0.16
34.4823
2
(
2)
52
29
82
52
Weibull
0.44
32.4190
3
(
2)
35
8.9
56
18
Condie
et
al.
(
1983)
Mouse
Male
Centrilobular
pallor
Gamma
0.38
30.0854
3
(
2)
44
7.4
61
15
Logistic
0.10
33.0715
2
(
2)
46
25
73
45
Log­
logistic
0.46
29.5813
3
(
2)
51
8.8
67
17
Log­
probit
0.46
29.5362
3
(
2)
56
20
70
28*

Multistage
0.31
30.6200
2
29
7.0
49
14
Probit
0.11
32.6482
2
(
2)
45
25
71
44
Weibull
0.35
30.4002
3
(
2)
32
7.2
50
15
Melnick
et
al.
(
1998)
Mouse
Female
Liver
hydropic
degeneration
Gamma
0.99
8.3184
3
(
1)
177
111
196
135
Logistic
1.0
10.2790
2
(
2)
190
85
199
123
Log­
logistic
1.0
8.2790
3
(
1)
191
123
199
146*

Log­
probit
1.0
10.2790
3
(
2)
189
127
198
146
Multistage
1.0
(?)
396.414
2
64
0.099
9.2
0.16
Probit
1.0
10.2790
2
(
2)
182
78
197
115
Weibull
1.0
10.2790
3
(
2)
172
88
196
118
NTP
(
1989a)
Mouse
Male
Stomach
nodules
Gamma
0.43
20.0036
3
(
2)
165
46
208
82
Logistic
0.22
22.0653
2
(
2)
158
77
223
131
Log­
logistic
0.48
19.6930
3
(
2)
165
55
206
89
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
58
Draft
­
Do
not
cite
or
quote
February
20,
2002
Log­
probit
0.50
19.4922
3
(
2)
176
66
211
95
Multistage
055
18.7358
1
116
32
167
66*

Probit
0.25
21.5725
2
(
2)
162
74
222
127
Weibull
0.38
20.6462
3
(
2)
137
35
188
68
CANDIDATE
STUDIES
FOR
THE
LONGER­
TERM
HA
Chu
et
al.
(
1982b)
Rat
Female
Hepatic
lesions
Gamma
flat
curve
fit
no
BMD
­­
­­
­­

Logistic
0.06
53.0088
2
(
2)
38
23
69
44
Log­
logistic
0.07
51.4919
3
(
2)
10
4.1
21
8.6
Log­
probit
0.05
52.6957
3
(
2)
36
18
51
26
Multistage
0.07
51.9025
2
16
8.2
32
17
Probit
0.06
52.9040
2
(
2)
35
22
65
42
Weibull
0.07
51.9025
3
(
2)
16
8.2
32
17
Chu
et
al.(
1982b)
Rat
Male
Liver
lesions
Gamma
0.45
56.7134
3
(
3)
31
1.6
36
3.3
Logistic
0.66
54.7588
2
(
2)
5.1
2.8
10
5.6
Log­
logistic
0.45
56.7134
3
(
3)
45
0.8
48
1.6
Log­
probit
0.45
56.7134
3
(
3)
37
4.3
41
6.1
Multistage
<
0.01
202.7
2
2.5
1.1
3
1.9
Probit
0.67
54.6647
2
(
2)
5.3
3.0
10
5.9*

Weibull
0.45
56.7024
3
(
3)
13
1.6
19
3.3
NTP
(
1989a)
Mouse
Female
Hepatic
vacuolization
Gamma
0.73
30.3981
3
(
2)
69
35
82
51
Logistic
0.27
33.9813
2
(
2)
68
41
96
66
Log­
logistic
0.80
30.0241
3
(
2)
70
38
87
54
Log­
probit
0.85
29.6545
3
(
2)
74
41
88
55*

Multistage
0.62
31.6064
2
53
24
76
45
Probit
0.33
33.0847
2
(
2)
68
40
95
64
Model
G­
O­
F
p
value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
A
­
59
Draft
­
Do
not
cite
or
quote
February
20,
2002
Weibull
0.61
31.3715
3
(
2)
56
28
81
46
CANDIDATE
STUDIES
FOR
THE
RfD
NTP
(
1989a)
Rat
Male
Hepatic
vacuolation
­
Subchronic
Gamma
0.70
65.8698
3
(
2)
2.2
1.3
4.4
2.6
Logistic
0.65
66.1399
2
(
2)
3.9
2.4
7.3
4.7
Log­
logistic
0.49
68.4402
3
(
3)
2.5
0.45
4.4
0.94
Log­
probit
0.61
66.4811
3
(
2)
5.5
2.9
8.0
4.2
Multistage
­
661.889
6
2.1
0.0064
2.4
0.0084
Probit
0.65
66.0888
2
(
2)
3.9
2.6
7.8
5.3
Weibull
0.70
65.8698
3
(
2)
2.2
1.3
4.4
2.6*

NTP
(
1989a)
Rat
Male
Fatty
Changes
in
Liver
­
Chronic
Gamma
1.0
84.7983
3
(
3)
29
0.66
32
1.4
Logistic
0.89
82.8323
2
(
2)
1.9
1.2
3.8
2.4
Log­
logistic
1.0
82.7983
3
(
2)
51
0.016
53
0.033
Log­
probit
1.0
84.7983
3
(
3)
38
0.97
41
1.4
Multistage
0.99
82.7983
2
8.9
0.66
13
1.4*

Probit
0.98
82.7996
2
(
2)
2.2
1.6
4.5
3.2
Weibull
1.0
84.7983
3
(
3)
12
0.66
16
1.4
*
Selected
model
results
for
endpoint.

AIC
Akaike
Information
Criterion
BMD
Benchmark
Dose
BMDL
95%
lower
confidence
level
on
BMD
G­
O­
F
Goodness­
of­
Fit
HA
Health
Advisory
?
Results
questionable
on
the
basis
of
visual
inspection
or
probable
calculation
error
W
BMDS
gave
a
warning
message:
"
BMDL
computation
is
at
best
imprecise
for
these
data"
B
­
1
Draft
­
Do
not
cite
or
quote
February
20,
2002
APPENDIX
B
Appendix
B
contains
BMD
Modeling
Output
in
Electronic
Format
(
compact
disk)
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
1
APPENDIX
C
Determination
of
the
Relative
Source
Contribution
for
Dibromochloromethane
(
DBCM)
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
2
The
relative
source
contribution
(
RSC)
is
the
percentage
of
total
daily
exposure
that
is
attributable
to
tap
water
when
all
potential
sources
are
considered
(
e.
g.,
air,
food,
soil,
and
water).
Ideally,
the
RSC
is
determined
quantitatively
using
nationwide,
central
tendency
and/
or
high­
end
estimates
of
exposure
from
each
relevant
medium.
In
the
absence
of
such
data,
a
default
RSC
ranging
from
20%
to
80%
may
be
used.
The
RSC
used
in
the
current
and
previous
drinking
water
regulations
for
DBCM
is
80%.
This
value
was
determined
by
use
of
a
screening
level
approach
to
estimate
and
compare
exposure
from
various
sources.
Information
considered
for
DBCM
during
this
process
is
summarized
below.

The
initial
step
in
RSC
determination
is
problem
formulation,
including
identification
of
population(
s)
of
concern,
critical
health
effects,
and
relevant
exposure
sources
and
pathways.
The
occurrence
of
DBCM
in
tap
water
is
reasonably
well
documented.
Occurrence
is
widespread
as
a
result
of
disinfection
of
drinking
water,
resulting
in
broad
exposure
of
the
U.
S.
general
population.
For
chronic
exposure
to
DBCM,
the
most
sensitive
responses
in
animal
studies
are
histopathological
changes
in
the
liver.
There
is
no
evidence
that
children
or
the
fetus
are
more
sensitive
to
these
effects
than
are
adults.
Although
polymorphisms
in
metabolizing
enzymes
might
predispose
some
groups
to
greater
sensitivity
to
this
compound,
no
sensitive
subpopulations
have
yet
been
clearly
identified.
Therefore,
the
population
of
concern
for
exposure
to
DBCM
is
considered
to
be
the
U.
S.
general
population.

Production
and
use
of
DBCM
occur
mainly
on
a
limited
scale
in
the
United
States.
In
the
past,
brominated
trihalomethanes
have
been
used
in
pharmaceutical
manufacturing
and
chemical
synthesis,
as
ingredients
in
fire­
resistant
chemicals
and
gauge
fluids,
and
as
solvents
for
waxes,
greases,
resins,
and
oils
(
U.
S.
EPA,
1975).
However,
use
patterns
have
changed
over
time.
DBCM
is
now
reportedly
used
in
laboratory
quantities
only
(
ATSDR,
1990).
Thus,
releases
to
the
environment
are
not
anticipated
to
be
significant
on
a
nationwide
basis
when
compared
to
occurrence
in
disinfected
tap
water.

DBCM
has
been
detected
in
air
and
food
in
a
few
studies,
in
addition
to
its
presence
in
tap
water.
No
data
were
available
in
the
materials
reviewed
for
levels
of
DBCM
in
soil.
DBCM
is
expected
to
volatilize
readily
from
wet
or
dry
soil
surfaces
based
on
its
Henry's
Law
constant
and
vapor
pressure
(
U.
S.
EPA,
1987).
For
this
reason,
exposure
via
ingestion
of
soil
is
not
expected
to
be
a
significant
route
of
exposure.
Therefore,
water,
food,
and
air
are
considered
to
be
the
relevant
pathways
for
this
analysis.

Evaluation
of
Occurrence
Data
The
next
step
in
RSC
determination
is
to
judge
whether
or
not
adequate
data
exist
to
characterize
exposure
from
relevant
exposure
pathways.
Factors
to
be
considered
in
the
evaluation
of
data
adequacy
include
sample
size;
whether
the
data
represent
a
random
sample
and
are
representative
of
the
target
population;
acceptable
analytical
detection
limits;
statistical
distribution
of
the
data,
and
estimator
precision.
In
addition,
it
is
important
to
know
whether
the
data
are
representative
of
current
conditions.
The
available
occurrence
data
for
DBCM
in
water,
air,
food,
and
soil
are
summarized
in
Chapter
IV
of
this
document.
Relevant
information
from
that
chapter
is
also
presented
below.
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
3
Occurrence
Data
for
Water
Adequate
data
are
available
to
estimate
central
tendency
and
high­
end
values
for
exposure
to
DBCM
from
treated
surface
and
ground
water.
Numerous
studies
(
summarized
in
Chapter
IV
of
this
document)
have
examined
the
levels
of
DBCM
in
disinfected
water.
Of
these
studies,
the
Information
Collection
Rule
(
ICR)
data
(
U.
S.
EPA,
2001)
most
closely
met
the
requirements
for
sample
size,
geographic
representation,
reporting
of
analytical
limits,
and
relevance
to
current
conditions.
This
survey
examined
the
occurrence
of
brominated
trihalomethanes
in
public
water
supplies
(
PWSs)
serving
at
least
100,000
persons
as
required
by
the
Information
Collection
Rule
promulgated
by
U.
S.
EPA
in
May
of
1996
for
disinfectants
and
disinfection
byproducts
(
D/
DBPs).
The
rule
covered
both
surface
and
ground
water
systems.
Monitoring
data
were
collected
from
about
300
water
systems
operating
501
plants
over
the18­
month
period
between
July
1997
and
December
1998.
At
each
plant,
samples
were
collected
monthly
and
analyzed
for
a
variety
of
D/
DBPs
on
a
monthly
or
quarterly
basis.
DBCM
was
among
the
analytes
evaluated
quarterly
(
U.
S.
EPA,
2001).
Five
samples
were
taken
each
quarter
at
each
plant
 
one
of
the
finished
water
and
four
of
the
water
in
the
distribution
system.
Of
the
four
samples
from
the
distribution
system,
one
represented
a
sample
with
the
same
residence
time
as
a
finished
water
sample
held
for
a
specific
period
of
time,
two
represented
approximate
average
water
residence
times
in
the
system,
and
one
sample
was
taken
where
water
residence
time
in
the
system
is
the
longest.
For
each
plant
and
reporting
period,
EPA
compiled
several
summary
statistics.
The
Distribution
System
(
DS)
Average
value
is
the
average
of
the
four
distribution
system
samples.
The
DS
High
Value
is
the
highest
concentration
of
the
four
distribution
system
samples
collected
by
a
plant
in
a
given
quarter.
The
DS
High
Value
might
be
from
any
of
the
four
samples
and
could
vary
from
quarter
to
quarter
depending
on
which
sample
yielded
the
highest
concentrations
in
each
quarter
(
U.
S.
EPA,
2001a).
Table
C­
1
summarizes
the
results
of
all
six
of
the
quarterly
reporting
periods.
The
DS
average
and
90th
percentile
values
for
DBCM
in
surface
water
were
4.72
µ
g/
L
and
5.57
µ
g/
L,
respectively.
The
DS
average
and
90th
percentile
values
for
dibromochloromethane
in
groundwater
water
were
3.09
µ
g/
L
and
8.94
µ
g/
L,
respectively.

U.
S.
EPA
set
a
minimum
reporting
level
(
MRL)
for
DBCM
of
1.0

g/
L
for
the
ICR.
The
MRL
is
a
level
below
which
systems
were
not
required
to
report
their
monitoring
results,
even
if
there
were
detectable
levels.
Values
below
the
MRL
were
assigned
a
value
of
zero
for
the
purpose
of
calculating
averages;
this
assignment
affects
the
calculation
of
mean
values
for
finished
water
and
DS
high
results
and
calculation
of
all
DS
average
values.

Data
for
Occurrence
in
Air
Occurrence
data
for
DBCM
in
ambient
outdoor
air
were
available
from
three
reports
(
Brodzinsky
and
Singh,
1983;
Shikiya
et
al.,
1984;
Atlas
and
Shauffler,
1991).
Brodzinsky
and
Singh
(
1983)
reviewed
and
summarized
existing
data
for
DBCM
concentrations
in
ambient
outdoor
air
for
several
urban/
suburban
or
source
dominated
locations
across
the
United
States
(
Table
C­
2).
No
concentration
data
were
available
for
rural
or
remote
areas.
The
authors
reported
mean,
median,
first
and
third
quartile
values,
and
minimum
and
maximum
values
by
city.
In
addition,
they
reported
the
same
measures
when
the
data
were
grouped
by
type
of
location
(
i.
e.,
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
4
Table
C­
1
DBCM
Concentrations
Measured
in
U.
S.
Public
Drinking
Water
Systems
Serving
100,000
or
More
Persons
Source
Data
Typea
Number
of
Samples
Medianb
Meanb
90th
Percentile
Range
DBCM
(

g/
L)

Surface
Water
Finished
1853
1.9
4.03
12.0
<
1.0
­
55.1
DS
Average
1655
2.40
4.72
13.2
0
­
67.3
DS
High
1655
2.9
5.57
15.0
<
1.0
­
67.3
Ground
Water
Finished
604
<
1.0
1.38
4.10
<
1.0
­
33
DS
Average
602
1.35
3.09
8.94
0
­
37.5
DS
High
602
2.1
4.60
12.9
<
1.0
­
85
Source:
Disinfectants
and
Disinfection
Byproducts
(
D/
DBPs)
ICR
Data,
U.
S.
EPA
(
2001).
a
Finished
=
sample
location
after
treatment,
before
entering
the
distribution
system
(
DS);
DS
Average
=
average
of
four
sample
locations
in
the
DS;
DS
High
=
the
highest
concentration
of
the
four
distribution
system
samples
collected
by
a
plant
in
a
given
quarter.
For
purposes
of
calculations,
all
values
below
the
minimum
reporting
level
(
MRL)
of
1.0

g/
L
for
all
three
compounds
were
assigned
a
value
of
zero.
b
Median
and
mean
of
all
samples
including
those
below
the
MRL.

urban/
suburban
or
source
dominated),
and
when
all
data
were
combined.
Dibromochloromethane
was
detected
in
the
air
samples
from
Magnolia,
AR,
El
Dorado,
TX,
Chapel
Hill,
NC,
Beaumont
TX,
and
Lake
Charles,
LA
at
mean
concentrations
of
0
ppt,
0.48
ppt,
14
ppt,
14
ppt,
and
19
ppt,
respectively.
Data
from
these
sites
were
combined
for
additional
statistical
analyses.
The
study
authors
indicated
that
a
value
of
0.0
was
entered
for
samples
below
the
detection
limit.
The
detection
limits
from
individual
studies
were
not
reported.
Mean
(
±
standard
deviation)
outdoor
air
concentrations
in
urban/
suburban
and
source
dominated
locations,
respectively,
were
15
±
4
ppt
and
0.28
±
0.67
ppt
for
DBCM.
Brodzinsky
and
Singh
(
1983)
also
calculated
overall
(
grand)
means
based
on
data
from
all
sites.
The
grand
mean
value
for
DBCM
was
3.8
ppt
(
n
=
89,
with
63
nondetects).
When
expressed
on
a
ng/
m3
basis,
the
corresponding
mean
value
was
0.032
µ
g/
m3.
Assuming
an
inhalation
rate
of
20
m3/
day,
this
concentration
results
in
a
daily
intake
of
0.6
µ
g/
day.
Assuming
a
rate
of
13.2
m3/
day,
this
concentration
results
in
a
daily
intake
of
0.43
µ
g/
day.

Shikiya
et
al.
(
1984)
analyzed
ambient
air
samples
collected
at
four
urban/
industrial
locations
in
the
California
South
Coast
Air
Basin
from
November
1982
to
December
1983
for
the
presence
of
DBCM.
The
sampling
locations
were
El
Monte,
downtown
Los
Angeles,
Dominguez,
and
Riverside.
The
air
samples
were
analyzed
using
gas
chromatography
with
detection
by
electron
capture.
The
quantitation
limit,
defined
as
a
level
10
times
greater
than
the
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
5
noise
level,
was
10
ppt
by
volume.
The
detection
limit
was
defined
as
three
times
the
noise
level.
Most
data
in
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
6
Table
C­
2
Selected
Concentration
Data
for
Individual
Brominated
Trihalomethanes
(
ppt)
in
Outdoor
Air
as
Summarized
in
Brodzinsky
and
Singh
(
1983)
a,
b
City
n
Nondetects
Mean
(
Std
dev.)
Median
3rd
Quartil
e
Maximum
Reference
DBCM
Individual
Sites
Beaumont,
TX
11
0
14
(
0.0)
14
14
14
Wallace
(
1981)

Chapel
Hill,
NC
6
0
14
(
0.0)
14
14
14
Wallace
(
1981)

El
Dorado,
AR
40
35
0.48
(
0.82)
0.0
0.82
2.5
Pellizzari
et
al.
(
1978)

Lake
Charles,
LA
4
0
19
(
9.6)
21
27
27
Pellizzari
(
1979)

Magnolia,
AR
28
28
0.0
(
0.0)
0.0
0.0
0.0
Pellizzari
et
al.
(
1978)

Totals
Urban/
Suburban
21
0
15
(
4.2)
14
14
27
­

Source
Areas
68
63
0.28
(
0.67)
0.0
0.0
2.5
­

Grand
Totals
89
63
3.8
(
6.7)
0.0
2.5
27
­

a
Includes
only
data
considered
to
be
of
adequate,
good,
or
excellent
quality
by
the
study
authors.
b
Concentrations
are
reported
as
parts
per
trillion
by
volume
this
report
were
presented
graphically.
A
few
additional
details
were
presented
in
a
short
summary
statement
for
each
chemical.
Summary
data
for
each
compound
included
monthly
means
and
composite
means.
The
monthly
means
were
calculated
as
the
average
of
all
data
at
a
site
that
were
above
the
quantitation
limit
for
a
single
month;
samples
with
concentrations
below
the
limit
of
detection
were
not
included
in
the
calculations.
The
composite
means
were
calculated
as
the
average
value
of
all
data
for
each
compound
above
the
quantitation
limit
at
each
site.
Only
seventeen
percent
of
the
samples
had
DBCM
levels
above
the
quantitation
limit
of
10
ppt
(
0.085
µ
g/
m3).
The
highest
reported
concentration,
monthly
mean,
and
mean
composite
for
DBCM
were
290
ppt
(
2.5
µ
g/
m3),
280
ppt
(
2.4
µ
g/
m3),
and
50
ppt
(
0.43
µ
g/
m3),
respectively;
all
were
recorded
in
downtown
Los
Angeles
in
June.
Only
two
monthly
means
were
above
160
ppt;
the
remainder
of
the
monthly
means
were
below
60
ppt.

Atlas
and
Schauffler
(
1991)
collected
replicate
air
samples
at
various
locations
on
the
Island
of
Hawaii
during
a
month­
long
field
experiment
to
test
an
analytical
method
for
determining
halocarbons
in
ambient
air.
DBCM
was
found
at
a
mean
level
of
0.27
ppt.
This
information
was
obtained
from
a
secondary
source
which
did
not
report
the
detection
limit.
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
7
Wallace
et
al.
(
1982)
conducted
a
pilot
study
designed
to
field
test
personal
air­
quality
monitoring
methods.
Personal
air
samples
were
collected
from
students
at
two
universities:
Lamar
University,
Texas,
located
near
a
petrochemical
manufacturing
area,
and
the
University
of
North
Carolina
(
UNC),
located
in
a
nonindustrialized
area.
The
samples
were
analyzed
for
a
number
of
volatile
organic
compounds,
including
brominated
trihalomethanes.
DBCM
was
not
detected
at
either
location.
Based
on
an
analytical
limit
of
0.12
µ
g/
m3
or
0.018
ppb,
these
data
suggest
that
exposure
via
personal
air
is
less
than
2.4
µ
g/
day.

There
are
several
limitations
associated
with
the
available
data
on
occurrence
of
DBCM
in
outdoor
air.
The
available
studies
were
collectively
limited
to
five
states
(
Arkansas,
California,
Hawaii,
North
Carolina,
and
Texas).
With
the
possible
exception
of
the
Hawaiian
study
(
Atlas
and
Schauffler
1991),
all
data
were
collected
from
urban/
suburban
or
source
dominated
locations.
Thus,
the
data
from
these
studies
are
not
considered
to
be
geographically
representative
of
the
United
States.
In
addition,
sample
size
was
not
explicitly
reported
in
the
Shikiya
et
al.
(
1984)
study
and
the
reported
means
were
based
only
on
data
above
the
detection
limit
(
only
17%
of
total
samples).
An
independent
statistical
evaluation
could
not
be
performed
because
raw
data
were
not
presented.
The
data
presented
by
Brodzinsky
and
Singh
(
1983)
were
obtained
from
multiple
sources
and
combined
results
for
sampling
periods
ranging
from
instantaneous
grab
samples
to
24
hour
averages
(
Wallace,
1997).
The
data
from
the
Shikiya
et
al.
(
1984)
and
Brodzinsky
and
Singh
(
1983)
reports
are
approximately
20
years
old
and
may
not
accurately
reflect
current
conditions.

Relatively
few
studies
have
reported
the
concentrations
of
trihalomethanes
in
indoor
air
of
homes.
Kostiainen
(
1995)
identified
over
200
volatile
organic
compounds
in
indoor
air
of
26
houses
identified
by
residents
as
causing
symptoms
such
as
headache,
nausea,
irritation
of
the
eyes,
drowsiness,
and
fatigue.
DBCM
was
not
reported
among
the
detected
compounds.

Weisel
et
al.
(
1999)
measured
brominated
trihalomethane
concentrations
in
indoor
air
in
49
New
Jersey
residences
selected
to
represent
low
and
high
levels
of
drinking
water
contamination
with
trihalomethanes.
Descriptive
statistics
for
DBCM
concentration
in
water
were
provided
for
the
combined
high
and
low
concentration
groups,
but
not
for
the
individual
groups.
One
valid
15­
minute
air
sample
was
collected
at
each
of
48
residences.
The
indoor
air
concentrations
of
DBCM
averaged
0.44
±
0.95
µ
g/
m3
(
0.052
±
0.11
ppb)
and
0.53
±
0.84
µ
g/
m3
(
0.062
±
0.09
ppb)
in
the
low
and
high
water
concentration
residences
with
detection
frequencies
of
5/
25
and
7/
23,
respectively.
The
detection
limit
was
0.14
µ
g/
m3
(
C.
Weisel,
personal
communication).
It
was
not
clear
whether
the
averages
were
based
on
all
measured
samples
or
only
those
samples
that
were
above
the
detection
limit.
For
this
reason,
the
data
were
not
used
for
calculation
of
exposure
to
DBCM
from
indoor
air.

It
is
possible
that
DBCM
has
been
surveyed
in
studies
of
volatile
organic
compounds
in
air
and
not
reported
because
it
was
below
detection
limits.
This
has
been
suggested
by
Dr.
Joachim
Pleil
of
the
U.
S.
EPA
Office
of
Research
and
Development,
who
is
highly
experienced
in
air
monitoring
of
volatiles
including
trihalomethanes.
According
to
Dr.
Pleil,
the
analytical
methods
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
8
used
in
analysis
of
volatile
organic
compounds
are
sufficiently
sensitive
to
detect
DBCM
even
if
is
present
only
in
minute
quantities.
For
example,
a
survey
of
volatile
organic
compounds
in
indoor
air
by
Pleil
et
al.
(
1985)
using
EPA
Method
TO­
14
(
detection
limit
approximately
100
ppt
by
volume,
or
0.85
µ
g/
m3,
for
trihalomethanes)
would
certainly
have
detected
DBCM
had
it
been
present
(
J.
Pleil,
personal
communication).

To
accurately
estimate
total
daily
inhalation
exposures
from
indoor
and
outdoor
air,
the
following
data
needs
to
be
evaluated:
location
and
season,
the
time
spent
indoors
compared
with
outdoors,
potential
exposures
of
individuals
while
showering
or
bathing,
potential
exposure
from
volatilization
of
DBCM
during
other
household
activities
(
e.
g.,
use
of
dishwashers,
toilet
flushing),
exposures
of
individuals
who
spend
large
amounts
of
time
at
indoor
pools
or
in
hot
tubs,
and
potential
for
occupational
exposures
(
e.
g.,
for
laundromat
or
sewage
treatment
plant
workers).
The
existing
measurement
data
are
not
adequate
for
such
a
refined
analysis,
but
may
be
used
to
roughly
estimate
intake
from
outdoor
air.

Data
for
Occurrence
in
Food
Information
on
the
levels
of
DBCM
in
foods
and
beverages
is
limited.
Chlorine
is
used
in
food
production
for
applications
such
as
the
disinfection
of
chicken
in
poultry
plants
and
the
superchlorination
of
water
at
soda
and
beer
bottling
plants
(
Borum,
1991).
Therefore,
the
possibility
exists
for
contamination
of
food
from
chlorination
by­
products
in
foods
with
resulting
dietary
exposure.

Two
studies
have
reported
analyses
of
commercial
beverages
for
DBCM.
In
Italy,
Cocchioni
et
al.
(
1996)
analyzed
61
samples
of
different
commercially
prepared
beverages
and
94
samples
of
mineral
waters
for
volatile
organo­
halogenated
compounds.
Maximum
DBCM
concentrations
of
13.9
µ
g/
L
(
ppb)
were
found
in
prepared
beverages,
with
a
frequency
of
detection
of
43%
(
26/
61),
with
a
detection
limit
of
less
than
1
µ
g/
L
(
ppb).
McNeal
et
al.
(
1995)
examined
27
different
prepared
beverages
and
mineral
waters
in
the
United
States
for
DBCM
at
a
detection
limit
of
0.1
ng/
g
(
ppb).
DBCM
was
detected
at
1
ng/
g
(
ppb)
in
only
one
of
seven
types
of
mineral
and
sparkling
waters
examined.
DBCM
was
not
detected
in
any
of
5
flavored
noncarbonated
beverages
examined.
DBCM
was
detected
in
only
4
of
the
13
carbonated
soft
drinks
examined
at
levels
of
0.5
to
2
ng/
g
(
ppb).
DBCM
was
not
detected
in
either
of
the
two
types
of
beer
examined.

Two
studies
have
tested
for
DBCM
in
individual
food
items.
McNeal
et
al.
(
1995)
tested
several
types
of
food
products
and
water
from
canned
vegetables
in
the
United
States
for
DBCM.
DBCM
was
not
detected
in
any
of
the
samples.
The
foods
examined
included
two
types
of
canned
tomato
sauce,
canned
pizza
sauce,
canned
vegetable
juice,
vegetable
waters
from
two
types
of
canned
green
beans
and
one
type
of
sweet
corn,
duck
sauces,
beef
extract,
and
Lite
syrup
product.
Imaeda
et
al.
(
1994)
examined
bean
curd
commercially
available
in
Japan
for
trihalomethanes.
DBCM
was
not
found
in
any
of
ten
samples
analyzed
at
a
detection
limit
of
0.1
ppb.
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
9
Kroneld
and
Reunanen
(
1990)
analyzed
pasteurized
and
unpasteurized
cow's
milk
for
DBCM
content
in
a
study
conducted
in
Turku,
Finland.
DBCM
was
detected
in
only
one
sample
of
pasteurized
milk
at
5
µ
g/
L
(
ppb).
The
detection
limit
was
not
specified
and
information
sample
size
was
unavailable
in
the
secondary
source
that
reported
this
study
(
U.
S.
EPA,
1994).
DBCM
was
not
detected
in
unpasteurized
milk.
The
presence
of
the
DBCM
in
pasteurized
milk
may
have
resulted
from
the
use
of
chlorinated
water
during
processing.

Estimates
for
dietary
intake
of
DBCM
by
residents
of
the
United
States
were
not
identified
in
the
materials
reviewed
for
this
document.
Information
on
the
levels
in
U.
S.
foods
is
too
limited
to
independently
calculate
a
reliable
estimate.
However,
the
available
data
suggest
that
the
concentration
of
DBCM
in
foods
is
low.

Data
for
dietary
intake
of
DBCM
are
available
from
a
study
conducted
in
Japan.
Toyoda
et
al.
(
1990)
analyzed
the
dietary
intake
of
DBCM
by
30
housewives
in
Nagoya
and
Yokohama.
Duplicate
portions
of
daily
meals
were
collected
for
three
consecutive
days,
sampled
for
DBCM
and
analyzed
at
a
detection
limit
of
0.2
ppb.
The
amount
and
types
of
food
consumed
were
not
reported.
This
omission
prevents
a
comparison
of
the
studied
diet
to
that
consumed
by
the
U.
S.
population.
The
concentration
of
DBCM
in
the
Japanese
diet
ranged
from
undetectable
to
0.6
ppb
(
average,
0.1
±
0.2
ppb),
and
the
mean
dietary
intake
was
estimated
to
be
0.3
±
0.3
µ
g/
day.
These
data
are
considered
adequate
only
for
a
rough
estimate
of
the
dietary
intake
of
DBCM.

Evaluation
The
occurrence
data
base
for
DBCM
in
tap
water
consists
of
nationally
aggregated
data
and
is
considered
adequate
for
determination
of
the
RSC.
In
comparison,
fewer
occurrence
data
are
available
for
DBCM
in
food
and
outdoor
air.
The
available
air
and
food
occurrence
data,
although
limited,
permit
rough
estimates
of
intake.

Determination
of
the
RSC
The
RSC
is
calculated
as
follows:

RSC
=
DI
water
(
1)
DI
total
where:

DI
water
=
DI
w
ater,
ingestion
+
DI
water,
inhalation
+
DI
water,
dermal
(
2)

DI
total
=
DI
water,
total
+
DI
outdoor
air
+
DI
food
(
3)

The
estimation
of
individual
terms
in
these
equations
is
described
below.

Exposure
Associated
with
Tap
Water
Uses
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
10
Exposure
to
DBCM
as
a
chlorination
by­
product
in
residential
water
can
occur
via
three
primary
exposure
routes:
1)
by
ingestion;
2)
by
inhalation
of
DBCM
volatilized
during
use
of
tap
water
for
bathing,
showering,
and
other
household
activities;
and
3)
by
dermal
exposure
during
showering
and
bathing.
The
existence
of
these
routes
for
DBCM
is
supported
by
recent
study
data.
Kerger
et
al.
(
2000)
demonstrated
that
levels
of
DBCM
in
indoor
air
are
related
to
the
use
of
tap
water
for
showering
and
bathing.
Increases
in
the
level
of
DBCM
in
the
breath
or
blood
after
showering
or
bathing
have
been
documented
in
human
subjects
(
Weisel
et
al.,
1999;
Backer
et
al.,
2000;
Lynberg
et
al.,
2001).
Quantitative
estimates
of
average
daily
exposure
from
volatilized
DBCM
or
dermal
contact
have
been
calculated
and
described
in
the
following
pages
for
comparison
with
other
routes
of
intake.
These
derived
values
have
solid
scientific
support
and
are
sufficient
for
a
reliable
estimate.
It
is
important
to
note
that
these
estimates
are
for
exposure
of
the
general
population
via
tap
water.
Individuals
who
participate
in
activities
such
as
swimming
or
hot
tub
use
may
experience
increased
dermal
and
or
inhalation
uptake
or
brominated
trihalomethanes
as
a
result
of
increased
contact
time
with
disinfected
water.
It
is
important
to
note
that
water
in
hot
tubs
and
swimming
pools
is
routinely
subjected
to
additional
disinfection
and
may
not
be
representative
of
tap
water
using
for
drinking,
cooking,
and
other
household
activities.

a.
Ingestion
of
DBCM
in
Drinking
Water
Ingestion
of
DBCM
is
calculated
by
multiplying
an
appropriate
intake
rate
for
tap
water
by
the
concentration
of
the
compound
found
in
tap
water.
Mean
water
intake
rates
of
1.2
and
0.6
L/
day
(
NRC,
1999)
were
used
for
total
mean
ingestion
from
all
uses
and
for
direct
ingestion
(
i.
e.,
direct
ingestion
from
tap,
does
not
include
use
of
tap
water
for
making
coffee
and
tea,
soup,
etc.),
respectively.
An
adjustment
for
intake
of
commercial
beverages
(
e.
g.,
soft
drinks
or
mineral
waters)
was
not
applied,
because
the
intake
of
these
beverages
would
be
subtracted
from
the
daily
tap
water
intake
and
the
available
data
(
e.
g.,
McNeal,
1995)
suggest
that
the
level
of
DBCM
in
such
beverages
is
usually
less
than
or
similar
to
the
level
found
in
tap
water.
Assuming
a
tap
water
concentration
of
4.72
µ
g/
L
(
the
distribution
system
average
for
DBCM
in
treated
surface
water),
the
total
and
direct
intakes
of
DBCM
via
ingestion
of
tap
water
are
5.7
µ
g
and
2.8
µ
g,
respectively.
b.
Inhalation
of
Waterborne
DBCM
A
three­
compartment
model
approach
was
used
to
investigate
the
exposure
from
waterrelated
dibromochloromethane
in
indoor
air.
The
three­
compartment
model
employed
was
that
of
McKone
(
1987).
This
model
predicts
the
concentration
of
a
volatile
chemical
in
water
(
in
this
case,
DBCM)
in
each
of
three
compartments
of
a
house:
the
shower,
the
bathroom,
and
the
remainder
of
the
house.
The
three­
compartment
model
recognizes
that
most
household
water
uses
are
episodic
rather
than
continuous,
and
room
barriers
(
walls,
doors)
may
restrict
the
rapid
mixing
of
DBCM
released
into
air
in
one
location
with
whole­
house
air,
leading
to
occasional
high
levels
of
DBCM
in
some
rooms
(
especially
those
with
high
water
usage,
such
as
the
shower
or
laundry).
Because
concentrations
are
not
constant,
results
are
calculated
as
a
function
of
time
throughout
the
day.
Based
on
the
time­
and
compartment­
specific
concentration
values,
human
exposure
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
11
levels
in
each
compartment
can
then
be
calculated
based
on
an
assumed
pattern
of
human
occupancy
and
behavior
within
the
house.
McKone
(
1987)
estimated
the
source
term
for
the
release
of
VOCs
from
water
to
air
in
each
of
the
three
compartments
by
extrapolation
from
measurements
of
radon
release
using
the
VOC­
specific
Henry's
law
constant
and
the
liquid­
and
gas­
phase
diffusion
coefficients.
Basic
equations
and
inputs
to
the
model
are
provided
below:
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
12
(
)
(
)
(
)
(
)
V
dy
t
dt
S
t
Q
y
t
Q
y
t
i
i
i
ji
j
ij
i
=
+
 
 
 


(
)
(
)
(
)
S
t
WU
TE
F
t
i
i
i
VOC
=
,
 
 
,
 

 
0
 
0
 

 
0
 
 
 
 
 
 
 
 
 
 
 
 
,

TE
TE
K
A
K
A
TE
D
HD
D
HD
i
VOC
i
Rn
OL
VOC
OL
Rn
i
Rn
L
G
Rn
L
G
VOC
=
 
=
 
+
 






+
 






25
1
25
1
2
3
2
3
2
3
2
3
.

.
/
/

/
/

(
)
(
)
E
ED
y
t
OF
t
dt
inh
i
i
=
 

1
Dose
E
BR
inh
=
 
Transient
three­
compartment
model
based
on
transfer
efficiency
developed
by
McKone
(
1987)

Where:

A
interfacial
area
existing
between
water
and
air
(
cm2)
BR
breathing
rate
(
L/
min)
C
aqueous­
phase
concentration
(
mg/
L)
D
L
liquid­
phase
diffusion
coefficient
(
cm2/
sec)
D
G
gas­
phase
diffusion
coefficient
(
cm2/
sec)
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
13
ED
exposure
duration
(
min)
E
inh
inhalation
exposure
(
mg/
L,
except
for
radon,
which
is
in
pCi/
L)
F(
t,
 i
0,
 i
*)
function
=
1
when
time
t
is
between
 i
0
and
 i
*,
and
0
otherwise
H
Henry's
law
constant
(
mg/
L
air/
mg/
L
water)
K
OL
overall
mass­
transfer
coefficient
(
cm/
min)
K
OL
A
overall
interfacial
mass­
transfer
coefficient
(
L/
min)
OF
i(
t)
occupancy
factor
for
compartment
i
at
time
t
(
1
if
present,
0
if
absent)
Q
ij
ventilation
rate
from
compartment
i
to
compartment
j
(
L/
min)
S
i
emission
rate
from
source
in
compartment
i
(
mg/
min)
TE
i
j
transfer
efficiency
of
chemical
j
during
water
use
in
compartment
i
(
1)

 i
0
time
when
water
device
in
compartment
i
starts
(
min)

 i
*
time
when
water
device
in
compartment
i
ends
(
min)
V
i
volume
of
compartment
i
(
L)
WU
i(
 i
0,
 i
*)
volume
of
water
used
in
compartment
i
between
time
 i
0
and
 i
*
(
L)
y
i
gas­
phase
concentration
in
compartment
i
(
mg/
L)
Rn
radon
voc
volatile
organic
compound
The
following
properties
of
dibromochloromethane
at
20
°
C
were
used
as
inputs
to
the
model:

H
(
mg/
L
air/
mg/
L
water)
=
0.036
D
L
(
cm2/
s)
=
9.63
x
10­
6
D
G
(
cm2/
s)
=
8.24
x
10­
2
The
following
human
activity
and
water
use
patterns
were
assumed.
Four
people
living
in
the
house
each
take
an
8­
minute
shower
every
morning,
and
spend
12
minutes
in
the
bathroom
immediately
thereafter.
Each
person
spends
an
additional
20
minutes
in
the
bathroom
some
time
during
the
remainder
of
the
day.
The
second
person
taking
a
shower
is
selected
for
the
purpose
of
comparing
exposure
estimates.
The
person
being
modeled
is
assumed
to
spend
75%
their
time
indoors
by
assigning
an
average
daily
occupancy
factor
of
0.75.

The
following
specific
parameters
were
used
to
implement
the
calculations
of
the
threecompartment
model
developed
by
McKone
(
1987):

Variable
Description
Value
PNUM
Number
of
people
in
the
house
4
Person
Designated
person
for
exposure
calculation
2nd
showerer
V
s
Volume
of
shower
2000
L
V
b
Volume
of
bathroom
10000
L
V
a
Volume
of
main
house
400000
L
I
s
Volume
of
water
used
in
shower
350
L
Variable
Description
Value
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
14
I
b
Volume
of
water
used
in
bathroom
350
L
I
a
Volume
of
water
used
in
main
house
450
L
R
s
Shower
air
residence
time
3
min
R
b
Bathroom
air
residence
time
12
min
R
a
Main
house
air
residence
time
78
min
Q
sb
Ventilation
rate
from
shower
to
bathroom
40000
L/
hr
Q
bs
Ventilation
rate
from
bathroom
to
shower
40000
L/
hr
Q
ab
Ventilation
rate
from
main
house
to
bathroom
50000
L/
hr
Q
ba
Ventilation
rate
from
bathroom
to
main
house
45000
L/
hr
Q
bo
Ventilation
rate
from
bathroom
to
outside
5000
L/
hr
Q
ao
Ventilation
rate
from
main
house
to
outside
258000
L/
hr
SFR
Shower
flow
rate
11
L/
min
TE
s
Rn
TE
from
shower
to
air
for
radon
0.7
TE
b
Rn
TE
from
bathroom
to
air
for
radon
0.3
TE
a
Rn
TE
from
household
water
to
air
for
radon
0.66
t
s
Time
in
shower
8
min
t
b
Time
in
bathroom
after
shower
12
min
t
b'
Time
in
bathroom
during
rest
of
day
20
min
t
s
0
Time
when
first
shower
water
use
starts
7:
00
a.
m.
t
s
*

Time
when
first
shower
water
use
ends
7:
08
a.
m.
t
b
0
Time
when
toilet
water
use
starts
12:
00
a.
m.
t
b
*

Time
when
toilet
water
use
ends
12:
00
a.
m.
t
a
0
Time
when
other
household
water
use
starts
7:
00
a.
m.
t
a
*

Time
when
other
household
water
use
ends
11:
00
p.
m.
OF
Daily
average
occupancy
factor
0.75
BR
Breathing
rate
9.2
L/
min
Source
for
BR
is
U.
S.
EPA
(
1995)
and
for
all
other
values
is
U.
S.
EPA
(
1993).

Based
on
these
assumptions
and
parameter
values,
the
model
predicts
an
inhaled
dose
for
DBCM
of
540
µ
g/
year
per
µ
g/
L
water..
Dividing
this
number
by
365
days/
year
and
multiplying
the
result
by
4.72
µ
g/
L
(
the
distribution
system
average
for
DBCM
in
treated
surface
water)
gives
an
estimate
of
7
µ
g/
day
for
inhalation
intake
of
DBCM
volatilized
from
tap
water.

c.
Dermal
Absorption
of
DBCM
via
Bathing
or
Showering
Estimates
of
dermal
uptake
of
DBCM
were
obtained
using
a
membrane
model
approach
as
described
in
Cleek
and
Bunge
(
1993)
and
Bunge
and
McDougal
(
1998).
The
approach
assumes
that
the
skin
is
composed
of
stratum
corneum
and
viable
epidermis.
If
the
concentration
of
the
vehicle
remains
constant
and
the
systemic
concentration
remains
small,
the
dermal
absorption
of
chemicals
can
be
divided
into
two
periods:
non­
steady
state
and
steady
state.
In
the
non­
steady
state
period,
the
chemical
is
absorbed
in
the
lipophilic
stratum
corneum.
The
viable
dermis
acts
like
a
sink
for
the
chemical
once
the
steady­
state
is
achieved.
Equations
are
available
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
15
DA
AC
K
K
L
t
event
w
p
sc
w
sc
event
=
2
/

 
for
calculation
of
uptake
under
both
steady
state
and
non­
steady
state
conditions.
Because
the
duration
of
exposure
to
DBCM
during
showering
and
bathing
(
10
minutes;
U.
S.
EPA,
1992)
is
substantially
less
than
the
time
to
steady
state
for
DBCM
absorption
through
skin
(
3.9
hours;
U.
S.
EPA;
1992),
the
non­
steady
state
approach
is
the
default
procedure
for
estimating
uptake:

(
4)

where:

DA
event
=
the
amount
of
DBCM
(
µ
g)
absorbed
in
a
10­
minute
shower
or
bath
A
=
the
surface
area
exposed
during
the
bath
or
shower
event
=
20,000
cm2
(
U.
S.
EPA,
1992)
C
w
=
concentration
of
DBCM
in
tap
water
=
4.72
x
10­
3
µ
g/
cm3
(
U.
S.
EPA,
2001)
K
p
=
permeability
coefficient
=
3.9
x
10­
3
cm/
hr
(
U.
S.
EPA,
1992)
K
sc/
w
=
stratum
corneum­
water
partition
coefficient
=
38
(
see
equation
5)
L
sc
=
diffusion
length
of
the
stratum
corneum
=
1.5
x
10­
3
cm
(
Bunge,
personal
communication)
t
event
=
duration
of
exposure
event
=
10
minutes
=
0.1667
hours
(
U.
S.
EPA,
1992)

The
stratum
corneum­
water
partition
coefficient,
Ksc/
w,
was
estimated
as
recommended
by
Bunge
and
McDougal
(
1998):

(
5)
K
K
sc
w
ow
/
.

=
=
0
71
38
where
K
ow
=
octanol
water
partition
coefficient
=
170
for
DBCM
(
U.
S.
EPA,
1992)

When
calculated
using
equation
(
4)
and
the
input
values
above,
the
dermal
uptake
of
DBCM
during
a
10­
minute
showering
or
bathing
event
is
approximately
0.65
µ
g.

Perhaps
the
most
difficult
part
of
estimating
dermal
exposure
is
the
determination
of
the
permeability
coefficient,
K
p,
in
the
equation
above.
Estimates
have
been
obtained
from
skin
penetration
experiments.
Most
skin
penetration
experiments
fall
into
one
of
two
categories:
in
vivo
experiments
performed
on
living
humans
or
animals,
and
in
vitro
experiments
made
in
diffusion
cells
with
excised
skin
from
humans
and
animals.
Determination
of
permeability
coefficients
in
vitro
and
in
vivo
generally
requires
that
the
exposure
concentrations
and
surface
area
are
known
and
consistent.
There
is
scientific
debate
over
whether
in
vitro
or
in
vivo
measurements
are
the
most
appropriate
way
to
measure
absorption
of
chemicals.
In
vitro
methods
can
provide
quick
and
direct
measures
of
flux
and
permeability
coefficients.
It
is
also
advantageous
that
human
skin
can
be
used
in
vitro
when
chemicals
would
be
too
toxic
in
in
vivo
studies.
Although
the
actual
in
vitro
experiments
are
simpler,
their
use
includes
many
important
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
16
log
.
.
(
log
)
.
K
K
MW
p
ow
=
 
+
 
2
72
0
71
0
006
log
.
.
(
log
)
.
K
K
MW
p
ow
hc
halo
=
 
+
 
2
72
071
0006
 
 
variables
(
e.
g.
animal
species,
thickness
of
skin,
use
of
fresh
or
frozen
skin,
and
receptor
solution)
and
uncertainties
which
influence
the
representativeness
of
the
data.
In
vivo
studies
are
often
more
elaborate
and
require
more
data
analysis.

The
U.
S.
EPA
(
1992)
Dermal
Exposure
Assessment
document
used
in
vitro
data
to
estimate
the
permeability
coefficient
(
K
p)
for
multiple
chemicals.
An
equation
developed
by
Potts
and
Guy
(
1993)
was
used
to
estimate
the
K
p
of
over
200
chemicals,
including
DBCM,
as
a
function
of
the
octanol:
water
partition
coefficient
(
K
ow)
and
molecular
weight
(
MW).
The
equation
was
derived
from
an
experimental
data
base
compiled
by
Gordon
Flynn
(
1990),
which
includes
data
for
in
vitro
dermal
absorption
of
about
90
chemicals
from
water:

(
6)

where,

log
K
ow
=
logarithm
of
the
octanol
water
partition
coefficient
=
2.23
for
DBCM
(
U.
S.
EPA,
1992)
MW
=
molecular
weight
=
208.28
for
DBCM
Despite
the
adequate
correlations
for
representing
experimental
permeability
data
for
a
broad
rage
of
chemicals,
experimental
data
may
deviate
from
predictions
made
using
the
Potts
and
Guy
equation
by
one
to
two
orders
of
magnitude.
This
variability
was
clearly
demonstrated
by
Vecchia
(
1997).
Although
uncertainty
in
experimental
temperature
and
other
data
are
partly
responsible,
other
known/
unknown
factors
may
also
contribute
to
this
discrepancy.
For
example,
the
correlation
assumes
that
MW
is
a
good
predictor
for
molecular
size.
This
assumption
may
not
be
appropriate
for
groups
of
compounds
with
chemical
diversities
affecting
molecular
size.
Halogenated
hydrocarbons
will
occupy
the
same
molar
volume
as
a
hydrocarbon
molecule
with
a
much
lower
MW.
As
a
result,
equations
based
on
MW
that
are
developed
from
databases
consisting
primarily
of
hydrocarbons
will
tend
to
systematically
underestimate
permeability
coefficients
for
chemically
dense
compounds
such
as
DBCM,
by
an
order
of
magnitude
or
perhaps
even
more
for
compounds
with
specific
gravity
values
larger
than
about
2.5
and
MW
greater
than
200.

The
K
p
was
calculated
using
the
procedure
of
Vecchia
and
Bunge
(
2002)
to
address
potential
underestimation
of
the
K
p
for
DBCM
by
the
prediction
method
used
in
U.
S.
EPA
(
1992).
This
calculation
used
a
modification
(
equation
7)
of
the
Potts
and
Guy
equation
(
equation
6)
which
incorporates
an
adjustment
factor
for
density
of
halogenated
compounds
when
compared
to
non­
halogenated
hydrocarbons:

(
7)

where:
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
17
log
K
ow
=
logarithm
of
the
octanol
water
partition
coefficient
=
2.23
for
DBCM
(
U.
S.
EPA,
1992)
MW
=
molecular
weight
=
208.28
for
DBCM
 hc
=
estimated
liquid
density
of
the
compounds
in
the
Flynn
equation
upon
which
equation
(
6)
was
developed
=
0.9
(
Vecchia
and
Bunge,
2002)

 halo
=
density
of
a
halogenated
hydrocarbon
=
2.38
for
DBCM
(
U.
S.
EPA,
1994)

The
resulting
value
for
K
p
is
0.03
cm/
hr.
Substitution
of
this
adjusted
value
in
equation
(
4)
results
in
an
uptake
estimate
of
2.0
µ
g
for
a
10­
minute
showering
or
bathing
event
This
value
is
about
3­
fold
higher
than
the
estimate
obtained
for
dermal
absorption
using
the
unadjusted
K
p
value
reported
in
U.
S.
EPA
(
1992).
The
2
µ
g
value
was
selected
for
determination
of
the
RSC
for
DCBM.

Exposure
Associated
Via
Food
and
Outdoor
Air
Dietary
intake
data
for
DBCM
from
a
Japanese
study
were
used
in
the
absence
of
intake
data
for
U.
S.
residents.
These
data
were
used
with
the
understanding
that
the
composition
(
and
thus
levels
of
DBCM)
of
U.
S.
and
Japanese
diets
may
differ.
The
grand
mean
for
outdoor
air
concentration
calculated
by
Brodzinsky
and
Singh
(
1983)
was
used
to
estimate
intake
from
outdoor
air.
Although
these
data
for
food
intake
and
air
concentrations
have
limitations,
they
were
considered
sufficient
for
a
screening
level
estimate
of
the
RSC.

Calculation
of
the
RSC
An
example
calculation
of
the
RSC
is
shown
below.
Results
of
calculations
using
different
exposure
assumptions
are
summarized
in
Table
C­
3.

DI
water
=
DIw
ater,
ingestion
+
DI
water,
inhalation
+
DI
water,
dermal
(
2)

=
5.7
µ
g/
day
+
7.0
ug/
day
+
2.0
µ
g/
day
=
14.7
µ
g/
day
DI
total
=
DI
air
+
DI
food
+
DI
water
(
3)

=
0.4
µ
g/
day
+
0.3
µ
g/
day
+
14.7
µ
g/
day
=
15.4
µ
g/
day
RSC
=
DI
water
=
14.7
µ
g
=
0.96
x
100
=
95%
DI
total
15.4
µ
g
where:

DI
water
=
Intake
of
DBCM
from
tap
water
DI
total
=
Intake
of
DBCM
from
all
relevant
sources
(
i.
e.,
water,
air,
and
food
for
DBCM)
DI
air
=
Intake
of
DBCM
from
outdoor
air,
assuming
0.032
µ
g/
m3
and
intake
of
13.2
m3/
day
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
18
DI
food
=
Intake
of
DBCM
from
food
=
0.3
µ
g/
day
after
Toyoda
(
1990)
DI
water,
ingestion
=
Intake
of
DBCM
from
tap
water
by
ingestion
(
assumes
intake
of
1.2
L/
day)
DI
water,
inhalation
=
Intake
of
DBCM
from
tap
water
by
inhalation,
as
determined
using
3­
compartment
model;
assumes
intake
of
13.2
m3/
day
(
U.
S.
EPA,
1995)
DI
water,
dermal
=
Dermal
absorption
of
DBCM
during
bathing
or
showering
using
adjusted
value
of
K
p;
assumes
1
shower
or
bath/
day
The
RSC
calculated
using
the
ICR
distribution
system
mean
for
surface
water
and
U.
S.
EPA
default
values
for
inhalation
and
drinking
water
ingestion
was
0.95
or
95%.
Substitution
of
groundwater
concentration
data
and/
or
the
direct
value
for
drinking
water
intake
also
resulted
in
RSC
values
greater
than
90%.
These
calculations
suggest
that
an
RSC
as
high
as
the
default
ceiling
of
80%
is
justified
for
DBCM.

The
uncertainty
in
use
of
80%
for
the
RSC
is
related
to
the
quality
of
data
for
intake
from
food
and
outdoor
air.
The
primary
concern
is
that
the
available
data
might
result
in
a
significant
underestimate
of
the
actual
exposure
via
these
media,
resulting
in
an
RSC
value
that
was
not
appropriately
protective
of
health.
A
series
of
calculations
was
performed
to
test
the
effect
of
underestimating
exposure
from
outdoor
air
and/
or
food
on
the
RSC.
An
arbitrary
10­
fold
increase
in
the
dietary
intake
of
DBCM
while
holding
intake
from
other
sources
constant
resulted
in
RSC
values
of
78%
or
81%,
depending
upon
the
intake
values
selected.
Increasing
the
intake
of
DBCM
from
outdoor
air
by
10­
fold
resulted
in
RSC
values
of
72%
or
76%.
Increasing
both
food
and
air
intake
of
DBCM
by
10­
fold
gave
RSC
values
of
62%
or
67%.
These
calculations
assumed
a
tap
water
concentration
of
4.72
µ
g/
L,
which
is
the
Information
Collection
Rule
distribution
system
mean
for
surface
water
(
U.
S.
EPA,
2001).
These
calculations
suggest
that
an
RSC
of
80%
would
be
protective
of
human
health
even
if
concentrations
of
DBCM
in
food
or
outdoor
air
were
underestimated
by
a
factor
of
almost
10­
fold.

Conclusion
An
RSC
of
80%
is
recommended
based
on
the
analysis
of
available
data
on
occurrence
of
DBCM
in
tap
water
and
other
media.
The
major
uncertainties
in
this
analysis
are
related
to
limited
measurement
data
for
DBCM
in
outdoor
air
and
in
food.
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
19
Table
C­
3
Results
of
RSC
Calculations
for
DBCM
Water
Source
Cw
(
µ
g/
L)
a
Condition
IRwater
(
L/
day)
b
IRair
(
m3/
day)
c
DIfood
(
µ
g/
day)
d
DIair
(
µ
g/
day)
e
DIwater
(
µ
g/
day)
f
DItotal
(
µ
g/
day)
g
RSC
(%)

Surface
4.72
Surface
Water
0.6
13.2
0.3
0.42
11.8
12.6
94
1.2
13.2
0.3
0.42
14.7
15.4
95
Ground
3.09
Ground
Water
0.6
13.2
0.3
0.42
10.9
11.6
94
1.2
13.2
0.3
0.42
12.7
13.4
95
Surface
4.72
If
intake
of
DBCM
from
food
increased
by
10­

fold
0.6
13.2
3.0
0.42
11.8
15.3
78
1.2
13.2
3.0
0.42
14.7
18.1
81
Surface
4.72
If
intake
of
DBCM
from
air
increased
by
10­
fold
0.6
13.2
0.3
4.2
11.8
16.4
72
1.2
13.2
0.3
4.2
14.7
19.2
76
Surface
4.72
If
intake
of
DBCM
from
air
and
food
increased
by
10­

fold
0.6
13.2
3.0
4.2
11.8
19.1
62
1.2
13.2
3.0
4.2
14.7
21.9
67
a
Concentration
of
DBCM
in
water:
ICR
distribution
system
mean
(
U.
S.
EPA,
2001).

b
Daily
intake
rate
for
water.
Values
are
for
direct
(
0.6
L/
day;
NRC.
1999)
or
total
mean
(
1.2
L/
day;
NRC,
1999)
ingestion
rates.

c
Daily
inhalation
rate.
Value
used
is
consistent
with
the
input
value
of
9.2
L/
min
for
the
three­
compartment
model
used
to
estimate
intake
of
DBCM
from
indoor
air.

d
Daily
intake
of
DBCM
in
food,
based
on
data
from
Toyoda
et
al.
(
1990).

e
Daily
intake
of
DBCM
in
air,
based
on
grand
mean
for
DBCM
concentration
in
outdoor
air
calculated
by
Brodzinsky
and
Singh
(
1983)

f
Daily
intake
of
DBCM
in
water
from
ingestion,
inhalation
of
volatilized
compound,
and
dermal
absorption.
See
text
for
details
of
calculation.

g
Total
daily
intake
of
DBCM
from
water,
outdoor
air,
and
food.
Draft
­
Do
not
cite
or
quote
February
20,
2002
C
­
20
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