United
States
Office
of
Science
Environmental
Protection
and
Technology
July
2003
Agency
Washington,
D.
C.

Office
of
Water
Drinking
Water
Criteria
Document
Brominated
Acetic
Acids
DRAFT
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
i
Draft,
do
not
cite
or
quote
Acknowledgements
Chemical
Manager/
Lead
Scientist:

Diana
Wong,
Ph.
D.,
DABT
(
Office
of
Water)

Contractor
Authors:

Lynne
Haber,
Ph.
D.
(
TERA)
Bonnie
Stern,
M.
P.
H.,
Ph.
D.
(
GRAM,
Inc.)
Claudine
Kasunic
(
GRAM,
Inc.)

EPA
Internal
Reviewer:

John
Lipscomb,
Ph.
D.,
DABT
(
NCEA/
ORD)
Linda
Teuschler,
Ph.
D.
(
NCEA/
ORD)

OST
Mail
Code
4304T
EPA­
822­
R­
03­
015
Title:
Drinking
Water
Criteria
Document
for
Brominated
Haloacetic
Acids:
External
Review
Draft
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
ii
Draft,
do
not
cite
or
quote
Table
of
Contents
List
of
Figures
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List
of
Tables
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v
I.
Executive
Summary
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I­
1
II.
Physical
and
Chemical
Properties
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II­
1
III.
Toxicokinetics
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III­
1
A.
Absorption
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III­
1
B.
Distribution
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III­
6
C.
Metabolism.
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III­
12
D.
Excretion
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III­
18
E.
Bioaccumulation
and
Retention
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III­
21
F.
Summary
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III­
24
IV.
Human
Exposure
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IV­
1
A.
Drinking
Water
Exposure
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IV­
1
A.
1
National
Occurrence
Data
for
MBA,
BCA,
and
DBA
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IV­
1
A.
0.1
ICR
Plants
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IV­
2
A.
0.2
Quarterly
Distribution
System
Average
and
Highest
Value
for
the
MBA,
BCA,
and
DBA
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IV­
2
A.
2
Factors
Affecting
the
Relative
Concentrations
of
MCA,
BCA,
and
DBA
in
Drinking
Water
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IV­
5
A.
2.1
Disinfection
Treatment
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IV­
6
A.
2.1.1
Disinfection
Treatment
in
ICR
Database
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IV­
9
A.
2.2
Bromide
Concentration
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IV­
15
A.
2.2.1
Bromide
Concentration
in
the
ICR
Database
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IV­
16
A.
2.3
Total
Organic
Carbon
(
TOC)
Concentration
in
ICR
Database
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IV­
22
A.
2.4
Seasonal
Shifts
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IV­
27
A.
2.4.1
Seasonal
Shifts
in
ICR
Database
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IV­
28
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
iii
Draft,
do
not
cite
or
quote
B.
Exposure
to
Sources
Other
Than
Drinking
Water
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IV­
32
C.
Overall
Exposure
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IV­
33
D.
Body
Burden
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IV­
33
E.
Summary
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IV­
34
V.
Health
Effects
in
Animals
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V­
1
A.
Short­
Term
Exposure
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V­
1
B.
Long­
Term
Exposure
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V­
19
C.
Reproductive
and
Developmental
Effects
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V­
22
D.
Mutagenicity
and
Genotoxicity
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V­
70
E.
Carcinogenicity
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V­
84
F.
Summary
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V­
85
VI.
Health
Effects
in
Humans
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VI­
1
VII.
Mechanisms
of
Toxicity
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VII­
1
A.
Mechanisms
of
Noncancer
Toxicity
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VII­
1
B.
Cancer
Mechanisms
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VII­
15
C.
Sensitive
Subpopulations
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VII­
16
D.
Interactions
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VII­
21
E.
Summary
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VII­
21
VIII.
Quantification
of
Toxicological
Effects
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VIII­
1
A.
Introduction
to
Methods
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VIII­
1
A.
2
Quantification
of
Noncarcinogenic
Effects
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VIII­
1
A.
1.1
Reference
Dose
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VIII­
1
A.
1.2
Drinking
Water
Equivalent
Level
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VIII­
4
A.
1.3
Health
Advisory
Values
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VIII­
5
A.
2
Quantification
of
Carcinogenic
Effects
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VIII­
7
B.
Noncarcinogenic
Effects
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VIII­
11
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
iv
Draft,
do
not
cite
or
quote
B.
1
Monobromoacetic
acid
.
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.
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VIII­
11
B.
1.1
One­
Day
Health
Advisory
for
MBA
.
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VIII­
12
B.
1.2
Ten­
Day
Health
Advisory
for
MBA
.
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VIII­
12
B.
1.3
Longer­
Term
Health
Advisory
for
MBA
.
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VIII­
13
B.
1.4
Reference
Dose
and
Drinking
Water
Equivalent
Level
for
MBA
.
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VIII­
14
B.
2
Bromochloroacetic
acid
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VIII­
14
B.
2.1
One­
Day
Health
Advisory
for
BCA
.
.
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VIII­
18
B.
2.2
Ten­
Day
Health
Advisory
for
BCA
.
.
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.
VIII­
18
B.
2.3
Longer­
Term
Health
Advisory
for
BCA
.
.
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.
.
.
VIII­
19
B.
2.4
Reference
Dose
and
Drinking
Water
Equivalent
for
BCA
.
.
.
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.
VIII­
19
B.
3
Dibromoacetic
acid
.
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.
VIII­
20
B.
3.1
One­
Day
Health
Advisory
for
DBA
.
.
.
.
.
.
.
.
.
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.
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.
.
VIII­
25
B.
3.2
Ten­
Day
Health
Advisory
for
DBA
.
.
.
.
.
.
.
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.
.
VIII­
26
B.
3.3
Longer­
Term
Health
Advisory
for
DBA.
.
.
.
.
.
.
.
.
.
.
.
VIII­
28
B.
3.4
Reference
Dose
and
Drinking
Water
Equivalent
Level
for
DBA
.
.
.
.
.
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.
VIII­
31
C.
Carcinogenic
Effects
.
.
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.
VIII­
32
C.
1
Monobromoacetic
acid
.
.
.
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.
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.
VIII­
32
C.
2
Bromochloroacetic
acid
.
.
.
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.
VIII­
33
C.
3
Dibromoacetic
acid
.
.
.
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VIII­
34
D.
Summary
.
.
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VIII­
36
IX.
References
.
.
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.
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.
.
.
.
.
.
IX­
1
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
v
Draft,
do
not
cite
or
quote
List
of
Figures
Figure
II­
1.
The
Chemical
Structures
of
MBA,
BCA,
and
DBA
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
II­
3
Figure
III­
1.
Proposed
Metabolism
of
DBA
.
.
.
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.
.
.
III­
14
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
vi
Draft,
do
not
cite
or
quote
List
of
Tables
Table
II­
1
Physical
and
Chemical
Properties
of
Brominated
Acetic
Acids
.
.
.
.
.
.
.
.
.
.
.
.
II­
3
Table
III­
1
Toxicokinetic
Data
for
BCA
and
DBA
in
F344
Rats
.
.
.
.
.
.
.
.
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.
.
III­
2
Table
IV­
1
Bromoacetic
Acids
Quarterly
Distribution
System
Average
and
Highest
Value
.
.
.
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.
IV­
4
Table
IV­
2
MBA
by
Disinfection
Method
.
.
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.
.
IV­
11
Table
IV­
3
BCA
by
Disinfection
Method
.
.
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.
.
IV­
12
Table
IV­
4
DBA
by
Disinfection
Method
.
.
.
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.
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.
.
IV­
13
Table
IV­
5
MBA
by
Influent
Bromide
Concentration
.
.
.
.
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.
.
IV­
18
Table
IV­
6
BCA
by
Influent
Bromide
Concentration
.
.
.
.
.
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.
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.
.
IV­
19
Table
IV­
7
DBA
by
Influent
Bromide
Concentration
.
.
.
.
.
.
.
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.
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.
.
IV­
20
Table
IV­
8
MBA
by
Influent
Total
Organic
Carbon
(
TOC)
Concentration
.
.
.
.
.
.
.
.
.
.
IV­
24
Table
IV­
9
BCA
by
Influent
Total
Organic
Carbon
(
TOC)
Concentration
.
.
.
.
.
.
.
.
.
.
.
IV­
25
Table
IV­
10
DBA
by
Influent
Total
Organic
Carbon
(
TOC)
Concentration
.
.
.
.
.
.
.
.
.
.
IV­
26
Table
IV­
11
MBA
by
Sample
Quarter
.
.
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.
IV­
29
Table
IV­
12
BCA
by
Sample
Quarter
.
.
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.
IV­
30
Table
IV­
13
DBA
by
Sample
Quarter
.
.
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.
IV­
31
Table
V­
1
Body
and
Liver
Weight
Changes
Induced
by
BCA
and
DBA
.
.
.
.
.
.
.
.
.
.
.
.
.
V­
8
Table
V­
2
Immunotoxicity
of
DBA
in
Female
B6C3F1
Mice
.
.
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.
.
V­
13
Table
V­
3
General
Toxicity
of
DBA
in
Female
B6C3F1
Mice
.
.
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.
.
V­
14
Table
V­
4
Reproductive
and
Developmental
Toxicity
of
BCA
Following
Peri­
conception
Exposure
.
.
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.
V­
28
Table
V­
5
Sperm
Quality
Parameters
in
Rats
Given
14
Daily
Doses
of
DBA
.
.
.
.
.
.
.
.
.
V­
41
Table
V­
6
Reproductive
Outcomes
in
Rats
Following
Oral
Dosing
with
DBA
.
.
.
.
.
.
.
.
V­
43
Table
V­
7
Outcome
of
Artificial
Insemination
of
Sperm
from
Rats
Dosed
with
DBA
.
.
.
V­
44
Table
V­
8
Reproductive
Organ
Weights
and
Sperm
Counts
in
Rats
Given
Daily
Doses
of
DBA
.
.
.
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.
.
V­
45
Table
V­
9
Sperm
Quality
Parameters
in
Rats
Given
Daily
Doses
of
DBA
.
.
.
.
.
.
.
.
.
.
.
V­
46
Table
V­
10
Average
Consumed
Daily
Doses
(
mg/
kg/
day)
for
Male
and
Female
Sprague­
Dawley
Rats
in
the
Two­
Generation
Reproductive/
Developmental
Toxicity
Study
.
.
.
.
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.
.
.
.
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.
.
.
.
V­
60
Table
V­
11
Incidences
of
Exposure­
Related
Histopathologic
Findings
in
the
Testes
of
Rats
Consuming
DBA
in
Drinking
Water
.
.
.
.
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.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
V­
62
Table
V­
12
Genotoxicity
Studies
of
MBA
.
.
.
.
.
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.
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.
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.
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.
.
.
.
V­
74
Table
V­
13
Genotoxicity
Studies
of
DBA
.
.
.
.
.
.
.
.
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.
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.
.
.
.
V­
83
Table
V­
14
Summary
of
Genotoxicity
Data
for
Brominated
Acetic
Acids
.
.
.
.
.
.
.
.
.
.
.
.
V­
84
Table
VIII­
1
Summary
of
Oral
Studies
of
MBA
Toxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII­
11
Table
VIII­
2
Summary
of
Oral
Studies
of
BCA
Toxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII­
15
Table
VIII­
3
Summary
of
Oral
Studies
of
DBA
Toxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
VIII­
20
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
1
Draft,
do
not
cite
or
quote
Chapter
I.
Executive
Summary
Three
brominated
acetic
acids,
monobromoacetic
acid
(
MBA),
dibromoacetic
acid
(
DBA),

and
bromochloroacetic
acid
(
BCA),
have
been
selected
for
evaluation
in
this
drinking
water
criteria
document
on
the
basis
of
(
1)
their
occurrence
in
drinking
water
as
chlorine
disinfection
byproducts,
and
(
2)
the
availability
of
toxicological
data
on
their
potential
human
health
effects.

MBA,
BCA,
and
DBA
are
water­
soluble
hygroscopic
crystals
in
pure
form,
and
are
very
soluble
in
water.
Brominated
acetic
acids
are
formed
during
ozonation
or
chlorination
of
water
that
contains
bromide
ions
and
organic
matter,
primarily
humic
and
fulvic
acids.
Formation
of
chlorinated
acetic
acids
is
higher
in
the
presence
of
humic
acid
fractions
of
water
than
in
the
presence
of
fulvic
acid,
suggesting
that
a
similar
relationship
may
hold
for
brominated
acetic
acids.

Bromide
ions
occur
naturally
in
surface
water
and
ground
water,
with
seasonal
fluctuations,
and
may
increase
due
to
saltwater
intrusions
under
conditions
of
drought
or
as
a
result
of
pollution.
In
the
presence
of
sufficient
concentrations
of
bromide
ion,
the
formation
of
brominated
compounds
may
be
favored
over
formation
of
chlorinated
compounds.
Brominated
acetic
acid
concentrations
in
drinking
water
are
typically
in
the
order
of
BCA>
DBA>
MBA.
No
toxicokinetic
studies
of
MBA
have
been
identified
in
the
literature.
Although
quantitative
information
on
BCA
and
DBA
toxicokinetics
is
limited
to
the
findings
in
a
single
comparative
toxicokinetic
study
with
rats,
the
data
demonstrate
that
both
compounds
are
rapidly
absorbed
from
the
gastrointestinal
tract,
almost
completely
metabolized,
and
minimally
excreted
in
the
urine
and
feces.
Following
a
single
intravenous
dose,
neither
BCA
nor
DBA
appeared
to
bind
significantly
to
plasma
proteins
or
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
2
Draft,
do
not
cite
or
quote
accumulate
in
blood
cells,
and
the
unbound
fraction
in
plasma
and
the
plasma:
blood
concentrations
was
close
to
unity.
Further,
the
apparent
volume
of
distribution
was
similar
to
the
total
body
water
volume
for
rats,
leading
the
authors
to
conclude
that
both
compounds
were
uniformly
distributed
outside
the
vascular
system
and
did
not
sequester
in
peripheral
tissues.

However,
in
the
absence
of
specific
tissue
measurements,
the
distribution
of
BCA
and
DBA
cannot
be
ascertained.
The
mechanisms
by
which
brominated
acetic
acids
are
metabolized
remains
unclear.
Potential
pathways
of
brominated
acetic
acid
metabolism
to
glyoxylic
acid
have
been
proposed
based
on
the
observed
metabolism
of
1,1,2,2­
tetrabromoethane,
and
based
on
analogy
to
chlorinated
acetic
acids
Metabolic
data
from
a
number
of
studies
demonstrate
that
chlorinated
acetic
acids
undergo
oxidative
dehalogenation
by
glutathione
transferase
zeta
(
GST­
Zeta)
activity
and
preliminary
data
indicate
that
a
similar
metabolic
pathway
is
likely
to
occur
for
the
brominated
acetic
acids.
It
is
not
clear
whether
the
toxicologically
effective
moiety
is
the
parent
compound
or
an
active
metabolite.
Both
BCA
and
DBA
are
rapidly
cleared
from
the
blood,
following
single
oral
or
intravenous
dosing,
although
these
data
are
inconsistent
with
the
results
of
repeatedexposure
drinking
water
studies.
Based
on
current
information,
brominated
acetic
acids
appear
to
be
rapidly
excreted
and
to
have
little
propensity
for
bioaccumulation.
DBA
administered
at
high
concentrations
to
pregnant
Sprague­
Dawley
females
was
reliably
measured
in
placental
tissue
and
in
fetal
plasma
at
concentrations
that
were
generally
similar
to
those
measured
in
maternal
plasma.

However,
quantifiable
levels
of
DBA
in
the
milk
of
the
lactating
rats
were
not
detected,
leading
to
the
conclusion
that
DBA
freely
crosses
the
placenta
and
distributes
to
the
fetus
during
gestation,

but
does
not
appear
to
bioaccumulate.
In
contrast,
preliminary
data
from
a
published
abstract
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
3
Draft,
do
not
cite
or
quote
reported
the
presence
of
DBA
in
the
milk
of
lactating
female
rats
at
concentrations
higher
than
those
in
blood
serum,
suggesting
to
the
authors
that
DBA
might
accumulate
in
milk.

EPA's
Information
Collection
Rule
(
ICR)
database
contains
extensive
information
on
concentrations
of
MBA,
BCA,
and
DBA
in
drinking­
water
systems,
and
on
how
those
concentrations
vary
with
input­
water
characteristics
and
treatment
methods.
The
database
contains
information
from
six
quarterly
samples
from
7/
97
to
12/
98,
from
approximately
300
large
systems
covering
approximately
500
plants.
The
mean
concentrations
of
BCA
were
1.47
and
3.61
µ
g/
L
from
groundwater
and
surface
water
respectively.
The
mean
concentrations
of
DBA
were
0.82
and
1.09
µ
g/
L
in
groundwater
and
surface
water,
respectively.
Statistical
analysis
of
these
data
indicated
that
the
mean
concentrations
of
MBA,
BCA,
and
DBA
in
surface
water
were
significantly
higher
than
the
mean
concentrations
of
these
chemicals
in
groundwater,
with
BCA
>

DBA
>
MBA
in
both
surface
water
and
groundwater.

The
concentrations
of
MBA
in
surface
water
treated
with
chlorine
were
similar
to
those
treated
with
chlorine
followed
by
chloramine.
BCA
and
DBA
concentrations
were
lower
when
free
chlorine
was
used
both
in
the
treatment
plant
and
the
distribution
system.
Although
ozonation
appeared
to
significantly
reduce
the
formation
of
BCA,
there
were
no
significant
differences
in
MBA
or
DBA
concentrations
with
the
use
of
ozone
in
treating
surface
water
as
compared
to
the
common
(
non­
ozonation)
chemical­
disinfection
processes.
In
addition
there
were
no
significant
differences
between
the
two
treatments
using
ozonation
in
treating
surface
water
for
MBA,
BCA
and
DBA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
4
Draft,
do
not
cite
or
quote
Consistent
with
the
findings
of
other
investigators,
and
the
chemistry
of
the
formation
of
bromoacetic
acids,
a
regression
analysis
of
the
ICR
data
indicated
that,
with
the
exception
of
MBA
in
surface
water,
there
was
a
significant
correlation
between
influent
bromide
concentration
and
the
mean
concentrations
of
BCA
and
DBA
in
surface
water
and
groundwater.

In
addition,
for
a
given
influent
bromide
concentration
range,
the
mean
concentrations
of
BCA
were
generally
significantly
higher
that
the
mean
concentrations
of
DBA
and
MBA
in
both
surface
water
and
groundwater.

A
regression
analysis
of
the
ICR
data
indicated
that
there
was
a
significant
correlation
between
influent
total
organic
carbon
(
TOC)
concentration
and
the
mean
concentrations
of
MBA,

BCA,
and
DBA
in
surface
water.
This
is
consistent
with
the
formation
of
brominated
acetic
acids
from
the
reaction
of
humic
acid
and
hypobromous
acid,
a
compound
formed
by
the
reaction
of
bromide
ion
with
ozone
and/
or
chlorine
in
the
disinfection
process.
In
addition,
for
a
given
influent
TOC
concentration
range
in
surface
water,
the
mean
concentrations
of
BCA
were
significantly
higher
than
the
mean
concentrations
of
DBA,
which
were
significantly
higher
than
the
MBA
mean
concentrations.

Based
on
only
two
seasons
of
monitoring,
statistical
analysis
indicated
that
the
mean
concentrations
of
MBA
in
surface
water
were
significantly
higher
in
the
summer
than
in
the
spring
Also,
based
on
only
two
seasons
of
monitoring,
the
mean
concentrations
of
BCA
in
surface
water
were
higher
in
summer
than
in
winter.
Aside
from
these
exceptions,
there
were
no
consistently
significant
differences
in
the
mean
concentrations
of
MBA,
BCA
or
DBA
between
one
season
and
another
in
either
surface
water
or
groundwater.
Seasonal
variations
in
brominated
acetic
acids
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
5
Draft,
do
not
cite
or
quote
may
be
dependent
on
seasonal
fluctuations
in
bromide­
ion
concentration,
which
were
not
monitored.

The
data
on
exposure
to
sources
other
than
drinking
water
are
limited,
but
MBA
has
been
used
in
industry
and
in
hospitals.
Between
1981
to
1983,
approximately
5000
workers
were
potentially
exposed
to
MBA..
No
data
were
located
on
exposure
to
MBA,
BCA,
or
DBA
in
food,

air,
or
via
dermal
exposure.
No
data
could
be
located
on
body
burden
levels
of
MBA,
BCA,
or
DBA.

The
available
toxicity
database
for
the
brominated
acetic
acids
is
limited
and
many
toxicity
endpoints
have
not
been
fully
explored.
However,
there
is
a
large
body
of
ongoing
work,

particularly
for
BCA
and
DBA.
Preliminary
results
for
many
studies
have
been
reported
in
published
abstracts
and
are
included
in
this
document
to
provide
a
sense
of
the
spectrum
of
effects
induced
by
the
brominated
acetic
acids.

Monobromoacetic
acid
The
toxicity
data
for
MBA
are
very
limited.
The
oral
LD
50
for
MBA
was
reported
as
177
mg/
kg
in
male
rats.
Oral
gavage
single­
dose
(
0
or
100
mg/
kg)
and
14­
day
studies
(
0
or
25
mg/
kg/
day)
have
been
conducted
to
assess
the
spermatotoxicity
of
MBA.
Neither
general
toxicity
nor
spermatotoxicity
were
observed
with
either
dosing
regimen.
In
a
published
abstract
on
the
developmental
toxicity
of
MBA,
decreased
maternal­
weight
gain,
decreased
live­
fetus
size,
and
increased
incidence
of
soft­
tissue
malformations
were
reported
at
gavage
doses
of

50
mg/
kg/
day
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
6
Draft,
do
not
cite
or
quote
administered
during
gestation
days
(
GD)
6­
15.
No
multi­
generation
reproductive
toxicity,

subchronic
or
chronic
studies
have
been
conducted
with
MBA.

MBA
is
a
dermal
irritant,
with
a
lowest­
observed­
effect
concentration
(
LOEC)
of
0.2
M
following
a
one­
hour
occluded
dermal
exposure
in
rabbits.
No
data
were
identified
for
the
toxicity
of
MBA
following
exposure
by
the
inhalation
route.

No
data
were
identified
on
the
carcinogenicity
of
MBA.
The
genotoxicity
data
for
MBA
have
provided
mixed
results.
MBA
was
mutagenic
in
Salmonella
typhimurium
and
induced
DNA
single­
strand
breaks
in
vitro,
but
did
not
induce
SOS
DNA
repair
(
a
DNA­
repair
system
induced
in
response
to
DNA
damage)
in
bacteria
or
micronuclei
in
a
newt­
larvae
system.

Bromochloroacetic
acid
Oral
toxicity
studies
of
BCA
have
identified
the
kidney,
liver,
and
developmental
organs
as
potential
targets
of
toxicity.
Increased
liver
weight
was
observed
at
the
highest
drinking
water
dose
tested
(
500
mg/
kg/
day)
in
B6C3F1
male
mouse
given
BCA
in
drinking
water
in
a
21­
day
study
evaluating
peroxisomal
proliferation
and
oxidative
damage.
Marginal
increases
in
liver
weight
were
induced
at
the
highest
dose
tested
(
39
mg/
kg/
day
in
drinking
water)
in
rats
evaluated
for
target­
organ
toxicity
as
part
of
a
26­
or
30­
day
reproductive
and
developmental
screening
assay.
In
this
assay,
treatment­
related
liver
histopathological
changes
(
cytoplasmic
vacuolization)

were
observed
beginning
at
5
mg/
kg/
day,
and
became
more
prominent
at
39
mg/
kg/
day
in
rats
given
BCA
for
30
days.
The
biological
significance
of
these
changes
was
unclear,
as
control
males
in
the
parallel
26­
day
study
exhibited
the
same
lesion
and
there
was
no
dose
response.
Overall,
the
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
7
Draft,
do
not
cite
or
quote
authors
concluded
that
the
high
dose
of
39
mg/
kg/
day
was
sufficient
to
cause
systemic
toxicity,

and
this
value
is
considered
to
be
a
marginal
LOAEL.

No
standard
developmental
toxicity
studies
have
been
conducted
with
BCA.
The
reproductive
toxicity
of
BCA
has
been
assessed
in
females
and
male
rats
exposed
to
BCA
in
drinking
water
for
30
to
34
days
during
the
peri­
conception
period,
which
included
a
12­
day
premating
period
of
exposure,
exposure
during
cohabitation
on
days
13­
18,
and/
or
gestational
exposures
of
varying
duration
(
e.
g.,
GD
1­
12,
GD
1­
21).
In
females
exposed
for
30
days,
the
NOAEL
for
reproductive
and
developmental
effects
(
decreased
live
fetuses/
litter
and
decreased
total
implants/
litter)
was
19
mg/
kg/
day,
and
the
NOAEL
for
maternal
toxicity
was
also
19
mg/
kg/
day,
based
on
kidney
toxicity
in
the
pregnant
dam.
In
females
exposed
only
from
GD
6
to
parturition,
no
dose­
dependent
increases
in
either
maternal
or
fetal
toxicity
were
observed.
No
effects
of
BCA
on
male
fertility
and
sperm
quality
were
noted.
In
another
study
(
reported
in
a
published
abstract),
male
Sprague­
Dawley
rats
administered
BCA
by
gavage
for
14
days
exhibited
impaired
sperm
motility,
abnormalities
in
sperm
morphology,
altered
spermiation,
and
reduced
fertility
(
evaluated
by
in
utero
insemination
of
untreated
females).
The
LOAEL
was
8
mg/
kg/
day,

and
a
NOAEL
could
not
be
determined.
Adverse
effects
on
sperm
quality
and
male
fertility
were
also
reported
(
in
a
published
abstract)
in
male
mice
exposed
to
BCA
for
14
days.
A
decrease
in
the
mean
number
of
litters
per
male
and
a
decrease
in
the
percent
of
live
litters
per
mated
female
were
observed,
with
a
NOAEL
of
24
mg/
kg/
day.
No
multigeneration
reproductive
toxicity
study
has
been
conducted
for
BCA.
However,
BCA
is
currently
undergoing
90­
day
subchronic
and
2­

year
chronic
bioassays.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
8
Draft,
do
not
cite
or
quote
QSAR
modeling
predicted
a
LOEC
for
skin
corrosion
of
0.7
M
for
BCA,
indicating
a
potential
for
dermal
irritation.
No
data
were
identified
for
the
toxicity
of
BCA
following
exposure
by
the
inhalation
route.

In
a
published
abstract,
BCA
was
reported
to
induces
liver
tumors
in
mice,
but
there
are
no
published
reports
of
a
full
bioassay
with
BCA.
BCA
was
mutagenic
in
Salmonella
typhimurium
and
induced
oxidative
DNA
damage,
as
measured
by
an
increase
in
the
DNA
adduct
8­
hydroxydeoxyguanisine
(
8­
OH­
dG),
in
the
livers
of
mice
given
BCA
in
drinking
water.
The
data
are
insufficient
to
evaluate
either
the
genotoxicity
or
the
potential
carcinogenicity
of
BCA.

Dibromoacetic
acid
In
a
range­
finding
study,
the
reproductive
and
developmental
toxicity
of
DBA,

administered
in
deionized
drinking
water
to
male
and
female
Sprague­
Dawley
rats,
was
evaluated.

Animals
were
exposed
beginning
14
days
prior
to
cohabitation
and
continuing
through
gestation
and
lactation
(
a
total
of
63
to
70
days
of
treatment).
Apparent
taste
aversion
was
associated
with
an
exposure­
dependent
reduction
in
water
consumption,
which
was
paralleled
by
a
reduction
in
food
intake
at
all
concentrations,
resulting
in
decreased
body
weights
in
parental
animals
and
postweanling
pups
at
the
two
highest
doses
tested.
The
only
observed
adverse
reproductive
effect
was
a
possible
reduction
in
mating
performance
in
the
highest
dose
group
(
1000
ppm),
as
evidenced
by
a
slight,
but
nonsignificant,
increase
in
the
number
of
days
of
cohabitation
and
a
decrease
in
the
number
of
mated
pairs
(
6/
10
in
the
1000
ppm
group
versus
9­
10/
10
in
all
other
groups).
No
effects
were
observed
on
the
incidence
of
pre­
and
post­
implantation
losses,
live
litter
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
9
Draft,
do
not
cite
or
quote
sizes,
and
gross
external
morphology
or
sex
ratios
in
the
pups.
Based
on
a
lack
of
reproductive
effects,
the
parental
and
reproductive/
developmental
NOAEL
for
this
study
was
1000
ppm.

In
a
recently
completed
a
two­
generation
reproductive
follow
up
study,
male
and
female
Sprague­
Dawley
rats
were
administered
DBA
in
drinking
water,
continuously
from
initiation
of
exposure
of
the
P
generation
through
weaning
of
the
F2
offspring.
Decreased
body
weight
gains
were
observed
in
high­
dose
P
males
and
females,
and
at
all
exposure
levels
for
F1
male
and
females,
attributed
to
a
general
retardation
in
growth
caused
by
decreased
water
and
food
consumption.
Observed
delays
in
sexual
maturation
in
the
F1
high­
dose
group
were
also
considered
to
be
due
to
growth
retardation.
No
adverse
treatment­
related
effects
were
observed
on
any
reproductive
index
or
developmental
parameters,
except
for
statistically
significant,

doserelated
increase
in
the
number
of
males
exhibiting
altered
spermatogenesis
(
i.
e.,
retained
Step
19
spermatids
in
Stage
IX
and
X
tubules,
and
increased
or
abnormal
residual
bodies
in
affected
seminiferous
tubules)
in
the
P
and
F1
groups,
and
testicular
malformation
in
four
males
in
the
high
dose
F1
group.
The
NOAEL
was
50
ppm
for
both
P
and
F1
generations.

A
series
of
rat
oral
gavage
studies
on
the
effects
of
DBA
on
spermatogenesis
and
male
fertility,
have
been
conducted
using
a
number
of
different
experimental
protocols,
including
a
single
high­
dose
study,
a
14­
day
study,
and
several
longer­
term
studies.
The
results
indicated
that
DBA
is
clearly
spermatotoxic,
as
demonstrated
by
histopathology
indicative
of
altered
spermiation.
In
the
14­
day
study,
the
LOAEL
was
10
mg/
kg/
day,
based
on
mild
histopathologic
changes
(
retention
of
Step
19
spermatids)
and
a
NOAEL
could
not
be
determined.
In
the
longerterm
studies,
male
rats
were
treated
for
up
to
79
days,
with
the
highest
dose
group
(
250
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
10
Draft,
do
not
cite
or
quote
mg/
kg/
day)
being
treated
for
only
42
days
due
to
the
onset
of
overt
toxicity.
However,
fertility
was
assessed
at
various
time
points
up
to
213
days
by
mating
treated
males
with
untreated
females.
The
severity
of
male
reproductive­
tract
toxicity
was
both
dose­
and
duration­
dependent.

In
the
group
of
males
given
250
mg/
kg/
day,
fertility
was
impaired
throughout
the
6­
month
recovery
period
following
cessation
of
treatment,
indicating
that
damage
was
structural
and
likely
permanent.
Mild
histopathologic
changes
(
retention
of
Step
19
spermatids)
were
observed
beginning
at
10
mg/
kg/
day
while
adverse
effects
on
sperm
quality
were
reported
beginning
at
50
mg/
kg/
day.
The
NOAEL
for
spermatotoxicity
was
2
mg/
kg/
day.
In
a
published
abstract,
rats
were
exposed
in
utero
from
GD
15
through
postnatal
day
(
PND)
98
to
DBA
in
drinking
water
and
reproductive
development
and
adult
reproductive
function
were
assessed.
Male
reproductive­
tract
development
(
as
indicated
by
delayed
preputial
separation),
as
well
as
spermatogenic
and
fertility
endpoints,
were
adversely
affected
at
the
lowest
dose
tested
(
50
mg/
kg/
day).
In
another
published
abstract,
exposure
of
male
rabbits
to
DBA
in
drinking
water,
for
a
period
beginning
in
utero
on
GD
15
and
continuing
to
24
weeks
of
age,
reduced
the
conception
rates
of
females
artificially
inseminated
with
sperm
from
these
treated
males.
The
LOAEL
was
0.97
mg/
kg/
day
and
a
NOAEL
could
not
be
determined.
Full
reports
of
these
studies
have
not
yet
been
published.

A
recently­
published
reproductive­
toxicity
study
did
not
demonstrate
significant
spermatotoxic
effects
in
male
rats
treated
with
single
oral­
gavage
doses
of
DBA,
as
evidenced
by
the
absence
of
treatment­
related
effects
on
sperm
motility,
morphology,
and
membrane
permeability
at
doses

600
mg/
kg,
although
mild
testes
histopathology
was
reported.
The
LOAEL
histopathology
was
600
mg/
kg
and
a
NOAEL
could
not
be
determined.
The
lack
of
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
11
Draft,
do
not
cite
or
quote
reported
spermatotoxicity
in
this
study
is
in
contrast
to
an
earlier
single­
dose
gavage
study,
in
which
significantly
adverse
effects
of
DBA
on
sperm
count,
motility,
and
morphology
at
doses

1250
mg/
kg
were
observed
in
a
different
strain
of
rats.
Variation
among
studies
in
DBA
spermatotoxic
effects
may
have
been
associated
with
differences
in
rat
strain,
experimental
design,

end­
point
measurements,
and
other
study
variables.

DBA
administered
to
Holtzman
rats
via
oral
gavage
on
GD
1­
8
did
not
impair
female
fertility
although
a
170
%
increase
in
serum
17 ­
estradiol
was
observed
at
250,
but
not
500,

mg/
kg/
day.
In
two
published
abstracts,
DBA
was
reported
to
induce
reproductive
and
developmental
toxicity
in
pregnant
CD­
1
mice
administered
DBA
on
GD
6­
15.
Delayed
parturition
was
observed
at
doses

24
mg/
kg/
day,
but
the
toxicological
significance
of
this
effect
is
unclear.
Increased
postnatal
mortality,
decreased
pup
weight,
and
tail
defects
were
observed
at

610
mg/
kg/
day.
In
the
second
abstract
by
the
same
authors,
the
ability
of
DBA
to
induce
fetal
malformations
was
examined
in
pregnant
CD­
1
mice
administered
DBA
by
oral
gavage
on
GD
6­

15.
The
NOAEL
for
renal
malformations
(
hydronephrosis)
was
50
mg/
kg/
day.
In
contrast,
no
treatment­
related
effects
on
litter
viability,
postnatal
mortality,
gross
malformations
and
a
wide
array
of
other
developmental
end
points
were
observed
in
a
recent
two­
generation
drinking
water
reproductive/
developmental
toxicity
study,
conducted
according
to
currently­
accepted
standard
test
guidelines.
Differences
in
findings
between
the
two­
generation
study
and
those
reported
in
published
abstracts
may
have
been
due
to
differences
in
internal
doses
associated
with
gavage
versus
drinking
water
DBA
administration,
species
differences
in
susceptibility
to
DBA
toxicity,

the
lower
mean
doses
tested
in
the
two­
generation
study,
and/
or
other
factors.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
12
Draft,
do
not
cite
or
quote
Among
the
three
brominated
acetic
acids,
MBA,
BCA,
and
DBA,
unequivocal
evidence
of
adverse
effects
on
spermatogenesis
in
rats
is
available
for
DBA.
Although
an
adverse
effect
on
the
mating
performance
in
male
rats
treated
with
high
doses
of
DBA
was
reported
,
reproductive
function
was
unaffected
by
altered
spermiation
in
the
two­
generation
reproductive/
developmental
toxicity
study.
In
the
single
study
identified
in
the
literature,
MBA
had
no
effect
on
spermatogenesis
under
treatment
conditions
similar
to
those
that
yielded
positive
indications
of
spermatotoxicity
for
DBA.
For
BCA,
no
effects
on
sperm
quality
were
reported
in
a
reproductive
and
developmental
toxicity
screening
assay.
Although
a
decrease
in
total
implants
per
litter
and
live
fetuses
per
litter
was
reported
for
BCA
in
this
study,
both
males
and
females
were
exposed
to
treated
drinking
water,
and
thus
these
reproductive
effects
might
have
been
due
to
female
exposure.
However,
adverse
effects
on
male
fertility
have
been
reported
in
two
other
studies
in
which
only
males
were
treated.

In
addition
to
reproductive
and
developmental
endpoints,
the
liver,
immunotoxicity,
and
neurotoxicity
of
DBA
have
been
evaluated.
In
male
mice
treated
with
DBA
in
drinking
water
for
21
days,
increased
liver
weight
was
observed
beginning
at
125
mg/
kg/
day
and
was
accompanied
by
oxidative
stress,
as
indicated
by
increases
in
hepatic
cyanide­
insensitive
Acyl­
CoA
activity
and
8­
OHdG
levels.
The
NOAEL
was
125
mg/
kg/
day.

In
an
immunotoxicity
assay
,
female
mice
were
given
DBA
in
their
drinking
water
for
28
days.
Four
independent
studies
were
conducted
and
different
endpoints
were
examined
in
each
study.
Studies
1­
3
investigated
selected
immunologic
parameters,
body­
weight
changes,
and
selected
organ
weights;
in
Study
4,
body
weight,
organ
weights,
hematology,
and
gross
pathology
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
13
Draft,
do
not
cite
or
quote
were
examined.
Overall,
the
results
of
Study
1
demonstrated
an
increase
in
several
measures
of
cellular
immunity,
including
an
increase
in
the
total
number
of
spleen
cells
and
an
increase
in
the
percent
of
spleen
cells
as
macrophages,
with
statistically
significant
effects
generally
occurring
at
doses

73
mg/
kg/
day.
However,
the
toxicological
significance
of
these
findings
was
unclear.
In
Study
2,
the
spleen
IgM
antibody­
forming
cell
response
was
significantly
decreased
at

70
mg/
kg/
day,
but
no
change
was
observed
in
serum
IgM
titer
(
a
more
generalized
measure
of
immune
function,
encompassing
immune
activity
in
the
bone
marrow
and
lymph
nodes
as
well
as
the
spleen).
No
change
in
macrophage
activation
was
observed
when
tested
for
in
Study
3.
In
Study
4,
the
authors
reported
decreased
body
weight;
decreased
thymus­
gland
weight;
increased
liver,
kidney,
and
spleen
weights;
and
increased
reticulocyte
counts.
For
this
group
of
studies,

spleen
IgM
antibody­
forming
cell
response
was
chosen
as
the
critical
effect
because
it
represented
a
clear
decrease
in
spleen
immune­
system
function.
The
NOAEL
for
this
end
point
was
38
mg/
kg/
day.
Changes
in
body
weight,
spleen
and
thymus
weights,
and
reticulocyte
counts
occurred
at
the
same
or
higher
doses
than
the
critical
effect;
changes
in
liver
and
kidney
weights
were
observed
at
lower
doses,
but
were
not
considered
to
be
toxicologically
significant
in
the
absence
of
supporting
clinical
chemistry
and
histopathology
data.

In
a
neurotoxicity
study
published
as
an
abstract,
male
and
female
adolescent
F344
rats
were
exposed
to
DBA
in
drinking
water
for
6
months,
and
a
neurobehavioral
test
battery
was
administered
to
all
animals
at
1,
2,
4,
and
6
months.
Dose­
dependent
neuromuscular
effects
included
mild
gait
abnormalities,
decreased
forelimb
and
hindlimb
strength,
hypotonia,
decreased
sensorimotor
responsiveness
(
as
measured
by
responses
to
a
tail
pinch
and
auditory
click),
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
14
Draft,
do
not
cite
or
quote
decreased
motor
activity,
and
a
chest­
clasping
response
that
was
only
observed
in
high­
dose
females.
Sensorimotor
responsiveness
did
not
progress
with
continued
exposure.
Neuropathologic
examination
showed
significant
myelin
fragmentation,
axonal
swelling,
and
axonal
degeneration
in
the
white
matter
of
the
spinal
cord,
and
eosinophilic
or
faintly
basophilic,
occasionally
vacuolated
swelling,
indicative
of
degenerating
axons,
in
the
spinal
cord
gray
matter.
Histological
evidence
of
neuropathology
was
observed
in
the
mid­
and
high­
dose,
and
was
not
evaluated
in
the
low­
dose
group.
The
LOAEL
for
this
study
was
20
mg/
kg/
day,
and
a
NOAEL
could
not
be
determined.

No
long­
term
systemic
toxicity
studies
for
any
exposure
route
were
identified
in
the
peerreviewed
literature.
However,
DBA
is
currently
undergoing
90­
day
subchronic
and
2­
year
chronic
bioassays.

Although
the
genotoxicity
database
is
limited,
DBA
is
mutagenic
in
Salmonella
typhimurium
assays
and
tests
for
DNA­
damage
repair,
and
has
increased
the
DNA
adduct,
8­

OHdG,
in
hepatic
DNA
of
mice
exposed
via
drinking
water..
On
the
other
hand,
no
induction
of
micronuclei
was
reported
in
a
newt­
larvae
system,
suggesting
that
DBA
was
not
clastogenic
in
the
newt
test
system.
The
clastogenicity
of
DBA
has
not
been
reported
in
other
assays
using
a
standard
test
protocol,
although
DBA
has
been
reported
to
be
co­
clastogenic.
A
standard
mousemicronucleus
assay
has
not
been
conducted.
These
data
support
the
conclusion
that
DBA
is
mutagenic
and
genotoxic,
although
the
nature
of
the
DNA
damage
induced
by
DBA
remains
unclear.
In
published
abstracts,
DBA
was
reported
to
induce
aberrant
crypt
foci
in
the
colon
of
rats
and
liver
tumors
in
mice.
However,
complete
reports
of
these
bioassays
have
not
been
published,
limiting
the
utility
of
these
data
in
assessing
the
potential
carcinogenicity
of
DBA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
15
Draft,
do
not
cite
or
quote
No
studies
were
identified
that
directly
evaluated
human­
health
effects
of
exposure
to
MBA,
BCA,
or
DBA
via
any
route
MBA
is
more
toxic
than
DBA
in
acute
toxicity
studies.
One
proposed
cellular
basis
for
the
toxicity
of
MBA
is
its
ability
to
inhibit
enzyme
activity
through
direct
alkylation
of
sulfhydryl
and
amino
groups.
This
hypothesis
is
supported
by
in
vitro
studies
using
purified
human
enzymes
and
by
evidence
for
DNA
alkylation,
but
a
direct
relationship
between
these
reactions
with
cellular
macromolecules
in
vivo
and
the
observed
toxic
effects
of
MBA
has
not
yet
been
established.

DBA
and
BCA
have
been
associated
with
liver,
kidney,
and
reproductive
and
developmental
toxicity
in
a
variety
of
toxicity
studies.
Potential
mechanisms
for
the
induction
of
adverse
liver
effects
include
perturbations
of
carbohydrate
homeostasis
or
toxicity
due
to
the
formation
of
reactive
metabolites
from
haloacetic
acid
or
tyrosine­
metabolism
pathways.
The
kidney
may
also
be
a
target
for
brominated
acetic
acids,
possibly
reflecting
direct
toxicity
associated
with
the
formation
of
reactive
metabolites,
or
toxicity
secondary
to
oxalate
formation
although
it
appears
unlikely
that
sufficient
oxalate
is
formed
during
brominated
acetic
acid
metabolism
to
adversely
affect
the
kidney.

Differences
in
the
spermatotoxicity
of
these
brominated
compounds
are
also
apparent.

DBA,
but
not
MBA,
induced
spermatotoxicity
at
gavage
doses
that
also
induced
overt
toxicity
in
an
acute
toxicity
study,
but
effects
on
reproductive
function
were
not
observed
in
a
twogeneration
reproductive
toxicity
drinking­
water
study
in
which
animals
were
administered
spermatotoxic
doses
of
DBA.
No
evidence
of
BCA­
induced
spermatotoxicity
was
found
in
a
reproductive
and
developmental
toxicity
screening
assay
.
However,
a
published
abstract
reported
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
16
Draft,
do
not
cite
or
quote
impaired
sperm
quality
and
spermiation,
and
reduced
male
fertility
as
assessed
by
in
utero
insemination
of
untreated
females
with
the
sperm
of
males
treated
with
BCA
for
14
days.
In
another
published
abstract,
BCA
was
also
reported
to
decrease
male
fertility
in
mice.
The
weightof
evidence
suggests
that
both
DBA
and
BCA
are
male
reproductive­
tract
toxicants.
One
hypothesized
target
for
observed
spermatotoxicity
is
the
Sertoli
cells.
Although
the
cellular
mechanisms
of
brominated
acetic
acid
spermatotoxicity
have
not
been
identified,
the
modification
of
key
proteins
necessary
for
Sertoli­
cell
function
or
direct
cytotoxicity
by
DBA
and/
or
its
reactive
metabolites
have
been
suggested
as
possible
mechanisms.
Another
potential
mechanism
of
spermatotoxicity
is
haloacetic
acid­
mediated
disruption
of
the
early
stages
of
steroidogenesis,

possibly
by
interfering
with
the
steroidogenic
acute
regulatory
protein
(
StAR)­
mediated
transport
of
cholesterol
within
the
mitochondrial
membrane
and
thereby
affecting
the
synthesis
of
pregnenolone,
the
precursor
of
progesterone.
Other
studies
suggest
that
brominated
acetic
acids
may
interfere
with
spermatogenesis
by
altering
sperm
proteins
(
most
notably
SP22)
that
play
an
important
role
in
the
fertilization
process,
possibly
by
regulating
the
androgen
receptor.
It
has
also
been
suggested
that
haloacetic
acids
acting
on
SP22
and
other
sperm
proteins
may
indirectly
compromise
androgen­
dependent
maintenance
of
spermatogenesis.

All
three
brominated
acetic
acids
have
been
reported
to
induce
developmental
effects.

Although
the
spectrum
of
developmental
endpoints
affected
by
in
vivo
treatment
does
not
implicate
a
common
mode
of
action,
the
results
of
whole­
embryo
culture
(
WEC)
testing
have
suggested
the
mechanisms
of
developmental
toxicity
among
haloacetic
acids
are
similar.
A
QSAR
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
17
Draft,
do
not
cite
or
quote
model
using
WEC
data
was
able
to
adequately
describe
the
rank­
order
potency
of
a
series
of
haloacetic
acids.
The
results
of
a
WEC
study
testing
mixtures
of
haloacetic
acids
were
consistent
with
the
QSAR
model
predictions
of
dose­
additivity.
Brominated
acetic
acids
also
induced
dysmorphogenesis
in
WEC
at
doses
lower
than
the
doses
of
known
metabolites,
suggesting
that
either
the
parent
compound
or
other
unidentified
metabolites
are
responsible
for
these
developmental
effects.
Apoptosis
induction
has
been
proposed
as
having
a
role
in
the
mechanism
of
onset
of
in
vivo
developmental
toxicity
based
on
the
results
of
WEC
testing,
but
this
hypothesis
has
not
been
confirmed
in
vivo
No
data
are
available
for
identifying
susceptible
populations.
In
addition,
no
data
on
agedependent
changes
in
the
expression
of
genes
involved
in
brominated
acetic
acid
were
located.

Based
on
the
results
of
some
in
vivo
developmental
toxicity
studies
reported
in
abstracts,
DBA,

but
not
MBA
or
BCA,
induced
fetal
toxicity
at
lower
doses
than
those
associated
with
maternal
effects,
suggesting
that,
at
least
for
DBA,
the
fetus
might
be
more
susceptible
than
the
adult.

However,
these
preliminary
developmental
studies
found
fetal
and
maternal
effects
only
at
doses
well
above
those
causing
effects
on
spermiation,
indicating
that
protection
against
the
latter
effect
would
also
provide
adequate
protection
to
children
and
fetuses.
Additionally,
DBA
was
administered
by
gavage
in
these
studies
and
observed
fetal
toxicity
might
have
been
routedependent
because
increased
fetal
susceptibility
is
not
supported
by
the
results
of
the
twogeneration
drinking
water
reproductive
toxicity
study
with
DBA.

There
are
also
inadequate
data
on
potentially
susceptible
subpopulations
based
on
genetic
variability.
Human
polymorphisms
in
GST­
Zeta,
the
enzyme
that
metabolizes
DBA
and
BCA
to
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
18
Draft,
do
not
cite
or
quote
glyoxylate,
have
been
characterized
by
several
investigators.
However,
in
the
absence
of
data
on
whether
the
parent
compound
or
a
metabolite
is
the
active
moiety,
the
functional
consequences
of
this
polymorphism
are
not
clear.
Individuals
having
underlying
defects
in
glycogen
storage
may
be
susceptible
to
liver
effects
induced
by
brominated
acetic
acids,
and
individuals
lacking
certain
enzymes
for
glyoxylate
metabolism
may
be
at
risk
for
BCA­
or
DBA­
induced
kidney
toxicity.
If
the
formation
of
reactive
oxygen
or
lipid
intermediates
is
responsible
for
the
toxicity
of
brominated
acetic
acids,
then
deficits
in
the
activity
of
anti­
oxidant
enzymes
might
also
represent
a
source
of
increased
susceptibility.
Another
potentially
susceptible
population
to
DBA
are
individuals
with
hereditary
tryosinemia
II
(
a
disease
involving
a
deficit
in
tyrosine
metabolism);
its
chlorinated
analog,
DCA,
has
been
shown
to
alter
tyrosine
metabolism
as
a
consequence
of
its
inhibitory
effects
on
GST­
Zeta.
None
of
the
possibilities
has
been
examined
directly
in
in
vivo
studies,
and
potentially
susceptible
populations
have
not
been
identified.

No
suitable
studies
for
the
derivation
of
drinking­
water
health
advisories
(
HA)
were
identified
in
the
literature
for
MBA,
BCA,
or
DBA.
Neither
subchronic
nor
chronic
toxicity
studies
have
been
conducted
with
any
of
these
compounds,
although
both
subchronic
and
chronic
toxicity
testing
of
BCA
and
DBA
is
planned
or
in
progress.
A
number
of
additional
studies
are
currently
ongoing.

There
are
no
human
epidemiology
studies
or
full
animal­
cancer
bioassays
for
MBA,
BCA,

or
DBA,
although
both
BCA
and
DBA
are
slated
for
full
testing
(
NTP,
2000b;
NTP,
2000c).

Under
the
1999
Draft
Guidelines
for
Cancer
Risk
Assessment,
the
data
are
inadequate
for
an
assessment
of
human
carcinogenic
potential
of
MBA
and
BCA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
I­
19
Draft,
do
not
cite
or
quote
There
is
concern
for
the
potential
carcinogenicity
of
DBA
based
on
preliminary
findings
reported
in
published
abstracts,,
and
analogy
to
DCA,
a
known
high­
dose
rodent­
liver
carcinogen.

However,
insufficient
data
are
available
to
assess
DBA
carcinogenic
hazard.
Under
the
1999
Proposed
Guidelines
for
Cancer
Risk
Assessment,
the
data
are
inadequate
for
an
assessment
of
human
carcinogenic
potential
of
DBA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
II­
1
Draft,
do
not
cite
or
quote
Chapter
II.
Physical
and
Chemical
Properties
Three
brominated
acetic
acids
have
been
selected
for
consideration
in
this
document.

These
are
monobromoacetic
acid
(
MBA),
dibromoacetic
acid
(
DBA),
and
bromochloroacetic
acid
(
BCA).
Brominated
acetic
acids
are
formed
during
ozonation
or
chlorination
of
water
containing
bromide
ions
(
Jacangelo
et
al.,
1989;
Pourmoghaddas
et
al.,
1993)
and
organic
matter,
primarily
humic
and
fulvic
acids.
Formation
of
chlorinated
acetic
acids
is
higher
in
the
presence
of
humic
acid
than
in
the
presence
of
fulvic
acid
(
WHO,
2000),
suggesting
that
a
similar
relationship
may
hold
for
brominated
acetic
acids.
Bromide
ions
occur
naturally
in
surface
and
ground
water.

However,
seasonal
fluctuations
in
bromide­
ion
levels
can
occur.
In
addition,
bromide­
ion
levels
can
increase
due
to
saltwater
intrusion
resulting
from
drought
conditions,
or
due
to
pollution
(
WHO,
2000).

The
bromide­
ion
concentration
is
an
important
determinant
of
the
spectrum
of
haloacetic
acids
formed
from
the
reaction
of
disinfectants
with
organic
material.
In
the
presence
of
sufficient
concentrations
of
bromide
ion,
the
formation
of
brominated
compounds
may
be
favored
over
formation
of
chlorinated
compounds
(
Pourmoghaddas
et
al.,
1993).
Brominated
acetic­
acid
concentrations
in
drinking
water
are
typically
in
the
order
of
BCA>
DBA>
MBA
(
Jacangelo
et
al.,

1989;
Krasner
et
al.,
1989;
Boorman
et
al.,
1999;
U.
S.
EPA,
2000a).
The
occurrence
of
these
compounds
is
discussed
more
fully
in
Chapter
IV.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
II­
2
Draft,
do
not
cite
or
quote
The
chemical
reactions
resulting
in
the
formation
of
brominated
acetic
acids
have
been
described
in
detail
for
chlorinated
water
(
WHO,
2000).
When
chlorine
gas
is
added
to
water
(
e.
g.,

as
a
disinfectant),
it
hydrolyzes
almost
immediately
to
form
hypochlorous
acid
(
HOCl):

Cl
2
+
H
2
O

HOCl
+
H+
+
Cl­

Hypochlorous
acid
can
then
dissociate
into
the
hydrogen
ion
and
hypochlorite
in
a
reversible
reaction:

HOCl

H+
+
OCl­

In
the
presence
of
bromide
ion,
hypobromous
acid
(
HOBr)
is
formed
from
hypochlorous
acid
in
the
following
irreversible
reaction:

HOCl
+
Br­

HOBr
+
Cl­

Similar
reactions
occur
to
form
HOBr
from
other
drinking­
water
disinfectants,
and
the
resulting
HOBr
reacts
with
organic
material
to
form
brominated
acetic
acids,
as
shown
below
(
Jacangelo
et
al.,
1989;
Pourmoghaddas
et
al.,
1993).
In
summary,
brominated
acetic
acids
are
formed
from
the
following
reactions:

Bromide
ion
+
Ozone
or
HOCl

HOBr
HOBr
+
Organic
acid
(
e.
g.,
humic
acid)

Brominated
acetic
acid
Figure
II­
1
shows
the
structure
of
monobromoacetic
acid
(
MBA),
bromochloroacetic
acid
(
BCA),
and
dibromoacetic
acid
(
DBA),
and
Table
II­
1
summarizes
key
physical
and
chemical
properties
of
these
compounds.
The
data
contained
in
Table
II­
1
apply
to
the
pure
form
of
the
selected
chemicals.
These
chemicals
exist
in
the
environment
in
a
dissolved
form.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
II­
3
Draft,
do
not
cite
or
quote
DBA
Br
C
OH
O
Br
H
C
MBA
Br
H
H
C
C
OH
O
BCA
Br
C
OH
O
Cl
H
C
Various
synonyms
exist
for
the
selected
chemicals
addressed
in
the
document.
Some
common
synonyms
for
MBA
are
bromoacetic
acid,
2­
bromoacetic
acid,

­
bromoacetic
acid,

bromoacetate
ion,
bromoethanoic
acid,
carboxymethyl
bromide,
and
acetic
acid,
bromo­.

Synonyms
for
DBA
are
dibromoacetate
and
acetic
acid,
dibromo­.
Likewise
for
BCA,
the
known
synonyms
are
acetic
acid,
bromochloro­,
and
chlorobromoacetic
acid.

Figure
II­
1.
The
Chemical
Structures
of
MBA,
BCA,
and
DBA
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
II­
4
Draft,
do
not
cite
or
quote
Table
II­
1.
Physical
and
Chemical
Properties
of
Brominated
Acetic
Acids.
a
Property
Monobromoacetic
acid
(
MBA)
Dibromoacetic
acid
(
DBA)
Bromochloroacetic
acid
(
BCA)

Chemical
Abstracts
Registry
Services
No.
79­
08­
3
631­
64­
1
5589­
96­
8
Formula
BrCH
2
COOH
Br
2
CHCOOH
BrClCHCOOH
Molecular
weight
138.95
217.84
173.39
Appearance
hygroscopic
crystal
hygroscopic
crystal
hygroscopic
crystal
Density
(
g/
mL)
1.93
 
1.98
Melting
point
(

C)
49­
51
49
27.5­
31.5
Boiling
point
(

C)
208
218
210­
215
Solubility
Water
Alcohol
Log
p
Log
De
miscible
miscible
0.41c
nd
very
soluble
very
soluble
1.22d
­
1.69
ndb
nd
1.08d
­
1.77
a.
Adapted
from
the
CRC
Handbook
of
Chemistry
and
Physics
(
1999),
The
Merck
Index
(
1996),
and
Sigma­
Aldrich
(
2000).

b.
nd:
no
data
c.
Log
p
is
the
value
derived
experimentally
as
presented
in
Hansch
et
al.
(
1995).

d.
Log
p
is
the
calculated
octanol
­
water
partition
coefficient
in
the
un­
ionized
form
as
presented
in
Schultz
et
al.
(
1999).

e.
Log
D
is
the
distribution
coefficient
between
n­
octanol
and
buffer
at
pH
7.4,
as
presented
in
Schultz
et
al.
(
1999).

PK
a
values
were
not
identified
for
these
compounds.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
1
Draft,
do
not
cite
or
quote
Chapter
III.
Toxicokinetics
Although
quantitative
toxicokinetic
data
for
brominated
acetic
acids
are
limited,
the
available
information
suggests
that
DBA
and
BCA
are
well
absorbed,
almost
completely
metabolized,
and
minimally
excreted
in
the
urine
and
feces.
Most
of
these
quantitative
data
come
from
a
single,
preliminary
study
conducted
by
Schultz
et
al.
(
1999).

A.
Absorption
Monobromoacetic
acid
No
studies
investigating
the
quantitative
parameters
of
MBA
absorption
were
identified
for
any
route
of
exposure.
Adverse
target­
organ
effects
observed
in
short­
term
toxicity
studies
(
described
in
detail
in
Section
V)
show
that
MBA
is
absorbed
following
exposure
by
the
oral
route;
however,
the
kinetics
of
absorption
are
not
currently
known.

Bromochloroacetic
acid
BCA
is
systemically
absorbed
following
oral
dosing
(
Table
III­
1).
Schultz
et
al.
(
1999)

administered
a
single
oral
gavage
or
intravenous
(
IV)
doses
of
500
µ
mol/
kg
(
87
mg/
kg
)
to
male
F344
rats.
Following
dosing,
BCA
venous
blood
concentrations
were
measured
at
0,
5,
10,
20,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
2
Draft,
do
not
cite
or
quote
Table
III­
1.
Toxicokinetic
Data
for
BCA
and
DBA
in
F344
Ratsa
Parameters
determined
following
IV
dosing
with
500
µ
mol/
kg
BCA
(
87
mg/
kg)
DBA
(
109
mg/
kg)

Area
under
blood
concentration­
time
curve
AUC
(
µ
M­
h)
576
±
286b
1120
±
362
Amount
excreted
in
urine
in
24
h
(%
Dose)
2.16
±
1.07
2.67
±
1.09
Steady­
state
apparent
volume
of
distribution
(
mL/
kg)
881
±
373
400
±
112
Total
body
clearance
(
mL/
h­
kg)
1037
±
453
491
±
116
Renal
clearance
(
mL/
h­
kg)
36.9
±
20.8
12.9
±
4.0
Mean
residence
time
(
h)
0.92
±
0.41
0.93
±
0.50
Elimination
half­
life
(
h)
3.93
±
1.50
0.72
±
0.12
Unbound
fraction
in
plasma
(
f
u)
0.93
±
0.07
0.89
±
0.05
Blood/
plasma
concentration
ratio
0.98
±
0.12
0.91
±
0.05
Parameters
determined
following
oral
dosing
with
500
µ
mol/
kg
BCA
(
87
mg/
kg)
DBA
(
109
mg/
kg)

Area
under
blood
concentration­
time
curve
AUC
(
µ
M­
h)
270
±
38
333
±
70
Mean
residence
time
(
h)
2.12
±
0.81
2.10
±
0.70
Time
to
peak
blood
concentration
(
h)
1.5
1
Mean
absorption
time
(
h)
c
1.20
1.17
Oral
Bioavailability
(%)
d
>
47e
>
30e
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
3
Draft,
do
not
cite
or
quote
Notes:
a.
Adapted
from
Schultz
et
al.,
1999
b.
Mean
±
standard
deviation
c.
Calculated
as
the
difference
between
the
mean
residence
time
following
IV
versus
oral
dosing.
d.
The
ratio
of
the
blood
concentration
AUC
for
oral
versus
IV
dosing
x
100%.
e.
Although
the
study
authors
reported
estimated
values
of
47%
and
30%
for
BCA
and
DBA,
respectively,
the
oral
bioavailability
is
likely
to
be
underestimated,
based
on
first­
pass
metabolism
considerations
associated
with
oral
and
not
IV
dosing.

30,
60
and
90
minutes,
and
3,
4,
6,
8
and
12
hours;
concentrations
of
BCA
in
the
urine
and
feces
were
measured
in
samples
collected
for
24
hours
after
dosing.
The
oral
bioavailability
­
the
ratio
of
the
averaged
values
for
the
area
under
the
curve
for
the
oral
and
i.
v.
doses
­
was
estimated
as
47%
for
BCA.
However
the
oral
bioavailability
of
BCA
might
be
higher
than
indicated
due
to
more
extensive
first
pass
metabolism
via
this
route
than
via
the
intraveneous
route.

To
measure
the
absorption
rate,
the
time­
to­
peak
blood
concentration
was
determined.

The
peak
concentration
of
BCA
was
observed
1.5
hours
following
oral
dosing.
Rapid
absorption
of
BCA
was
confirmed
by
the
mean
absorption
time,
which
was
determined
by
measuring
the
difference
in
the
mean
residence
time
in
blood
between
oral
and
IV
dosing.
The
mean
absorption
time
was
reported
as
1.2
hours.
These
data
show
that
BCA
is
readily
absorbed
following
a
single
bolus
dose.
However,
no
quantitative
data
are
available
to
assess
whether
the
degree
of
absorption
would
be
different
under
ingestion
conditions
more
closely
resembling
human
exposure
conditions
(
i.
e.,
temporally
dispersed
and
at
much
lower
doses).

No
studies
on
the
absorption
of
BCA
were
identified
following
exposure
by
the
inhalation
or
dermal
routes.

Dibromoacetic
acid
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
4
Draft,
do
not
cite
or
quote
Shultz
et
al.
(
1999)
also
studied
the
systemic
absorption
of
DBA
following
oral
dosing,
as
described
above
for
BCA
(
Table
III­
1).
Male
F344
rats
were
given
single
oral
gavage
or
IV
doses
of
500
µ
mol/
kg
(
109
mg/
kg)
DBA.
Following
dosing,
the
DBA
venous­
blood
concentrations
were
measured
at
0,
5,
10,
20,
30,
60
and
90
minutes,
and
4,
6,
and
8
hours;
concentrations
of
DBA
in
the
urine
and
feces
were
measured
in
samples
collected
24
hours
after
dosing.
The
oral
bioavailability
of
DBA
was
estimated
to
be
30%.
However,
this
value
is
likely
to
be
an
underestimate
because
it
probably
reflects
first
pass­
metabolism
following
oral
dosing
and,
thus,

underestimates
the
degree
of
oral
absorption.
The
peak
blood
concentration
was
observed
1
hour
following
oral
dosing
and
the
mean
absorption
time
(
the
difference
in
the
mean
residence
time
in
blood
following
dosing
via
oral
and
IV
routes)
was
1.17
hours,
indicating
rapid
absorption.

However,
no
quantitative
data
were
provided
on
the
degree
of
absorption
under
conditions
of
low­
dose
repeated
oral
dosing,
a
regimen
which
would
more
closely
resemble
human­
exposure
conditions.

Blood­
level
measurements
for
brominated
acetic
acids
following
oral
dosing
have
been
reported
in
other
toxicity
studies,
and
provide
some
additional
information
on
DBA
absorption.

As
part
of
a
study
on
the
effects
of
DBA
on
pubertal
development
and
adult
reproductive
function,
Klinefelter
et
al.
(
2000,
abstract
only)
reported
blood­
serum
and
milk
concentrations
in
Sprague­
Dawley
rats
(
3
litters/
dose)
administered
drinking
water
containing
0,
400,
600,
or
800
ppm
DBA
from
gestation
day
(
GD)
15
through
postnatal
day
(
PND)
98.
Estimated
doses
resulting
from
these
treatments
were
0,
50,
75,
and
100
mg/
kg/
day,
respectively
(
personal
communication).
DBA
levels
were
also
assayed
in
dams'
milk
and
blood
serum,
and
in
the
serum
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
5
Draft,
do
not
cite
or
quote
of
suckling
males
on
PND
20
(
personal
communication).
Only
data
for
the
800
ppm
treatment
group
were
presented
in
the
abstract.
In
this
group,
DBA
concentrations
were
5.2­
14.1
µ
g/
mL
in
dams'
milk,
3.0­
6.9
µ
g/
mL
in
dams'
serum,
and
0.01­
0.24
µ
g/
mL
in
the
serum
of
male
offspring.

The
limited
amount
of
information
in
the
abstract
precluded
quantification
of
the
degree
of
absorption.
However,
the
data
demonstrate
that
DBA
was
absorbed
following
drinking­
water
exposure.
In
contrast,
no
detectable
levels
of
DBA
were
observed
in
the
plasma
of
female
B6C3F1
mice
following
28
days
of
exposure
to
DBA­
treated
drinking
water
at
concentrations
up
to
2000
mg/
L,
corresponding
to
estimated
doses
of
229
to
285
mg/
kg/
day
depending
on
the
substudy
(
NTP,
1999).
The
reasons
for
the
differences
in
findings
between
the
Klinefelter
(
2000)
and
NTP
(
1999)
studies
are
unclear,
and
may
have
been
due
to
differences
in
analytical
methods.
No
details
on
the
post­
dosing
sampling
schedules
for
blood
and/
or
milk
were
presented
in
either
study.
The
absence
of
measurable
DBA
in
plasma
in
the
NTP
(
1999)
study
might
reflect
extensive
metabolism
and
rapid
excretion,
rather
than
limited
absorption.
Alternately,
species
differences
and/
or
differences
in
hormonal
status
might
have
affected
the
kinetics
of
DBA
absorption,

distribution,
metabolism
and/
or
excretion
in
these
studies.
Christian
et
al.
(
1999)
reported
the
detection
of
DBA
in
the
plasma
of
male
and
female
rats
exposed
to
DBA
in
drinking
water
concentrations
for
16
hours,
and
following
14
days
of
treatment,
also
demonstrating
that
gastrointestinal
absorption
occurred.

No
studies
on
the
absorption
of
DBA
were
identified
following
exposure
by
the
inhalation
or
dermal
routes.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
6
Draft,
do
not
cite
or
quote
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
7
Draft,
do
not
cite
or
quote
B.
Distribution
Monobromoacetic
acid
No
studies
of
MBA
tissue
distribution
following
dosing
by
any
route
were
identified.

Bromochloroacetic
acid
Schultz
et
al.
(
1999)
administered
male
F344
rats
single
oral
gavage
or
IV
doses
of
500
µ
mol/
kg
(
87
mg/
kg)
BCA.
BCA
venous­
blood
concentrations
were
measured
at
0,
5,
10,
20,
30,

60
and
90
minutes,
and
at
3,
4,
6,
8
and
12
hours
post­
dosing.
No
concentrations
in
tissues
other
than
blood
were
assessed.
Following
intravenous
dosing,
BCA
did
not
appear
to
bind
significantly
to
plasma
proteins
or
accumulate
in
blood
cells.
The
unbound
fraction
in
plasma
was
0.93,
and
the
ratio
between
plasma
and
blood
concentrations
was
close
to
unity
(
i.
e.,
0.98).
Tissue
concentrations
were
not
measured
directly.
However,
following
intravenous
dosing,
the
apparent
volume
of
distribution
(
881
mL/
kg)
was
similar
to
the
total
body­
water
volume
for
rats
(
approximately
660
mL/
kg)
(
Reinoso
et
al.,
1997).
This
similarity
suggested
to
the
study
authors
that
BCA
distributed
uniformly
outside
the
vascular
system
and
was
unlikely
to
sequester
significantly
in
peripheral
tissues.
The
octanol­
buffer
partition
coefficient
(
Log
D),
considered
to
be
a
reasonable
predictor
of
lipophilicity
(
Schultz
et
al.,
1999)
was
reported
to
be
­
1.77
at
pH
of
7.4.
This
low
value
also
suggested
that
at
physiological
pH
values,
BCA
has
little
propensity
to
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
8
Draft,
do
not
cite
or
quote
accumulate
in
fat
tissue.
However,
in
the
absence
of
direct
tissue
measurements,
the
distribution
of
BCA
cannot
be
ascertained.
Plasma
binding,
plasma:
blood
concentrations,
and
the
apparent
volume
of
distribution
(
V
ss)
were
only
measured
following
intravenous
dosing
and,
thus,
the
effects
of
first­
pass
metabolism
are
not
known.
Further,
the
high
dose
employed
in
this
study
might
have
resulted
in
metabolic
saturation
which
could
have
led
to
a
wider
distribution
of
BCA
than
would
have
occurred
at
lower
doses
where
metabolism
was
not
saturated.

Short­
term
animal
studies
suggest
that
oral
administration
of
high
doses
of
BCA
results
in
liver,
reproductive,
and
developmental
toxicity
(
NTP,
1998,
Parrish
et
al.,
1996;
see
Chapter
5
for
more
details),
indicating
that
BCA
does
distribute
to
the
liver,
the
reproductive
organs,
and
the
fetus
under
the
conditions
of
these
studies.
The
low
protein­
binding
capacity
of
BCA
suggests
that
the
potential
for
distribution
across
the
placenta
to
the
fetus
may
be
significant
at
high
maternal
doses
administered
during
pregnancy.

No
data
on
tissue
distribution
of
BCA
following
exposure
by
the
inhalation
or
dermal
routes
were
identified.

Dibromoacetic
acid
The
systemic
distribution
of
DBA
has
also
been
studied
by
Schultz
et
al.
(
1999).
Male
F344
rats
were
given
single
oral
gavage
or
IV
doses
of
500
µ
mol/
kg
(
109
mg/
kg)
DBA.
DBA
venous­
blood
concentrations
were
measured
at
0,
5,
10,
20,
30,
60
and
90
minutes,
and
at
4,
6,

and
8
hours
post­
dosing.
No
tissue
concentrations
other
than
blood
were
measured.
Following
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
9
Draft,
do
not
cite
or
quote
intravenous
dosing,
DBA
did
not
bind
significantly
to
plasma
proteins
or
accumulate
in
blood
cells.
The
unbound
fraction
in
plasma
was
0.89
for
DBA,
and
the
ratio
between
plasma
and
blood
levels
of
DBA
was
close
to
unity
(
i.
e.,
0.91).
Tissue
concentrations
were
not
measured
directly,

but
the
apparent
volume
of
distribution
and
the
total
body­
water
volume
in
rats
were
similar,

suggesting
to
the
study
authors
that
DBA
was
widely
and
uniformly
distributed
outside
the
vascular
system
and
was
unlikely
to
sequester
significantly
in
peripheral
tissues.
The
octanolbuffer
partition
coefficient
(
Log
D),
considered
to
be
a
reasonable
predictor
of
lipophilicity
(
Schultz
et
al.,
1999),
was
reported
to
be
­
1.69
at
pH
of
7.4.
This
low
value
suggested
that
at
physiological
pH
values,
DBA
has
little
propensity
to
accumulate
in
fat
tissue.
However,
in
the
absence
of
direct
tissue
measurements,
the
distribution
of
DBA
is
not
known.
Plasma
binding,

plasma:
blood
concentrations,
and
the
apparent
volume
of
distribution
(
V
ss)
were
only
measured
following
intravenous
dosing
and,
thus,
the
effects
of
first­
pass
metabolism
cannot
be
ascertained.

Further,
the
high
dose
employed
in
this
study
might
have
resulted
in
metabolic
saturation,
which
could
have
led
to
a
wider
distribution
of
DBA
than
would
have
occurred
at
lower
doses
where
metabolism
was
not
saturated.

Short­
term
animal
studies
suggest
that
oral
administration
of
DBA
results
in
toxicity
to
the
liver,
kidney,
spleen,
and
male
reproductive
system.(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b;

Linder
et
al.,
1995;
Parrish
et
al.,
1996;
Linder
et
al.,
1997;
Cummings
and
Hedge,
1998;
Vetter
et
al.,
1998;
NTP,
1999;
more
details
in
Chapter
5),
indicating
that
DBA
does
distribute
to
these
organs
under
the
dosing
conditions
of
these
studies.
Further,
the
study
by
Klinefelter
et
al.
(
2000,

abstract),
described
in
the
previous
section
on
DBA
absorption,
showed
that
the
milk
of
lactating
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
10
Draft,
do
not
cite
or
quote
Sprague­
Dawley
females
treated
with
800
ppm
(
100
mg/
kg/
day)
of
DBA
in
drinking
water
contained
elevated
levels
of
DBA
relative
to
the
females'
blood
serum
(
5.2­
14.1
µ
g/
mL
in
milk,

3.0­
6.9
µ
g/
mL
in
serum).
Thus,
lactational
distribution
to
nursing
pups
may
be
significant
at
high
maternal
doses.

As
part
of
a
series
of
range­
finding
reproductive
and
developmental
toxicity
studies,

Christian
et
al.
(
2001)
evaluated
the
distribution
of
DBA
in
adult
Sprague­
Dawley
rats,
and
in
pregnant
females
and
their
fetuses.
Male
and
female
rats
(
10/
sex/
group)
were
given
DBA
in
deionized
drinking
water
at
concentrations
of
0,
125,
250,
500
or
1000
ppm,
beginning
14
days
prior
to
cohabitation
and
continuing
through
gestation
and
lactation
(
63­
70
days
of
treatment).

The
average
daily
doses
(
based
on
measured
water
consumption
and
body
weights)
varied,

depending
on
the
phase
of
reproduction.
For
parental
males
throughout
the
study
(
SD
1­
70),

equivalent
mean
daily
doses
were
10.2,
20.4,
35.7,
and
66.1
mg/
kg/
day,
respectively.
For
females
on
SD
1­
15,
equivalent
mean
daily
doses
were
13.3,
26.2,
41.8
and
60.2
mg/
kg/
day,
respectively;

and
14.8,
30.3,
48.5
and
81.6
mg/
kg/
day,
respectively,
on
gestation
day
(
GD)
0­
21.
During
lactation
(
LD
1­
29),
the
estimated
doses
were
was
43.5,
86.6,
150.7
and
211.7
mg/
kg/
day
for
the
0,
125,
250,
500,
and
1000
ppm
groups,
respectively;
however,
these
doses
included
consumption
of
water
by
the
pups
and
thus
overestimated
the
mean
daily
intake
for
lactating
females.

An
additional
6
male
and
17
female
rats/
group
were
used
for
collecting
bioanalytical
samples.
Blood
plasma
levels
of
DBA
in
parental
male
and
female
rats
were
taken
on
SD
1
and
14,
in
mated
females
on
GD
20
and
LD
15,
and
in
weanling
male
and
female
rats
on
LD
29
(
which
was
also
postweaning
day
1).
Tissue
levels
of
DBA
in
placenta
and
amniotic
fluid
were
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
11
Draft,
do
not
cite
or
quote
assessed
on
GD
21
and
in
milk
on
LD
15.
During
the
designated
collection
days,
three
plasma
samples
were
collected
from
male
and
female
rats
at
approximately
8­
hour
intervals
in
order
to
assess
the
potential
differences
in
tissue
concentrations
associated
with
diurnal
rhythms
of
water
consumption
by
animals.

In
analyzing
biodisposition
samples,
numerous
values
were
measured
at
concentrations
below
the
limit
of
detection
(
LOD)
of
the
methodology
used.
These
values
were
quantified
but
were
not
considered
by
the
study
authors
to
be
reliable.
In
blood
plasma,
DBA
was
detected
in
male
rats
at
125
ppm
and
reliably
quantified
at

250
ppm
after
24
hours
of
continual
access
to
DBA
in
drinking
water.
Blood
plasma
levels
in
males
exposed
to
125
ppm
did
not
exceed
the
level
of
detection
even
after
14
days
of
exposure.
The
overall
pattern
of
DBA
detection
in
plasma
in
female
rats
was
similar
but
more
variable;
however,
quantifiable
DBA
was
increased
in
a
dosedependent
fashion
at
all
exposure
concentrations
on
SD
14,
and
was
observed
at
concentrations
of
4.4­
6.7
µ
g/
g
at
the
end
of
the
dark
period.
The
higher
plasma
levels
at
this
time
were
attributed
to
nocturnal
drinking
habits
and
not
to
accumulation
of
DBA
in
plasma.
During
gestation,
the
calculated
daily
dose
of
DBA
was
increased,
and
was
reflected
in
increased
plasma
concentrations
at
all
DBA
exposure
levels
on
GD
20
(
ranging
from
4.2­
18.0
µ
g/
g),
but
not
in
an
exposuredependent
manner.
During
lactation,
DBA
could
be
quantified
in
the
maternal
plasma
at
all
doses
when
measured
on
LD
15.
Measurements
ranged
from
1.9
to
26.2
µ
g/
g
and
varied,
depending
on
time
of
sample;
a
dose­
response
was
only
observed
at
one
time
point.
Analysis
of
milk
samples
collected
from
three
lactating
dams/
group
during
the
light
period
on
LD
15
indicated
that
DBA
was
not
detected
in
milk,
in
contrast
to
the
findings
of
Klinefelter
et
al
(
2000,
abstract).
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
12
Draft,
do
not
cite
or
quote
Pooled
fetal
blood,
amniotic
fluid,
and
placentas
were
collected
on
GD
21
from
three
litters/
exposure
group
and
analyzed
for
DBA.
In
the
placenta,
DBA
was
detected
at
all
exposure
levels
but
could
only
be
reliably
quantified
in
the
1000­
ppm
group
(
8.1
µ
g/
g
at
a
mean
maternal
daily
dose
of
8.16
mg/
kg/
day).
In
fetal
plasma,
DBA
was
detected
in
a
dose­
dependent
manner
in
all
groups;
however,
levels
in
the
125­
and
250­
ppm
groups
were
below
the
limit
of
detection
and
reliable
quantitation
was
only
possible
in
the
500
and
1000
ppm
groups
(
2.4
and
9.2
µ
g/
g
at
maternal
doses
of
48.5
and
81.6
mg/
kg/
day,
respectively).
In
contrast,
quantifiable
amounts
of
DBA
were
noted
in
the
amniotic
fluid
(
3.9,
5.3,
3.0,
and
5.8
µ
g/
g
at
14.8,
30.3,
48.5,
and
81.6
mg/
kg/
day,
respectively)
and
levels
were
comparable
to
those
observed
in
the
plasma
of
maternal
rats
on
GD
20
at
the
start
of
the
dark
period.
demonstrated
a
dose­
dependent
distribution
of
DBA
in
amniotic
fluid
(
3.9­
5.8
µ
g/
g)
on
GD
21
at
DBA
drinking
water
concentrations
of
125­
1000
ppm
(
estimated
mean
daily
doses
of
14.8­
81.6
mg/
kg/
day).
Equivalent
concentrations
of
serum
plasma
DBA
in
pregnant
females
on
GD
20,
taken
at
the
same
time
during
the
dark
cycle
as
the
amniotic
fluid
samples,
ranged
from
3.6
to
6.4
µ
g/
g.
These
results
demonstrate
that
at
high
drinking
water
concentrations,
DBA
can
cross
the
placenta
and
distribute
to
fetal
tissue.

No
data
on
tissue
distribution
of
DBA
following
exposure
by
the
inhalation
or
dermal
routes
were
identified.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
13
Draft,
do
not
cite
or
quote
C.
Metabolism
Monobromoacetic
acid
No
studies
were
identified
that
described
the
metabolism
of
MBA
following
exposure
by
any
route.

Bromochloroacetic
acid
Most
of
the
data
on
the
metabolism
of
BCA
are
indirect
or
based
on
analogy
to
the
chlorinated
acetic
acids.
These
indirect
data
suggest
that
dihalogenated
acetic
acids
are
rapidly
metabolized.
Schultz
et
al.
(
1999)
did
not
directly
measure
the
metabolism
of
BCA.
However,

comparisons
of
renal
and
blood
clearance
following
IV
administration
suggested
that
metabolism
was
likely
to
be
the
major
contributor
to
BCA
removal
from
the
blood.
Only
2%
of
BCA
blood
clearance
was
accounted
for
by
renal
clearance
and
excretion
in
the
feces
was
negligible.
Thus,

non­
renal
clearance
(
e.
g.,
through
metabolism)
accounted
for
most
of
the
removal
of
the
parent
compound
from
the
blood.
It
should
be
noted
that
these
data
were
obtained
following
IV
administration
of
the
test
compound
and,
thus,
the
degree
of
metabolism
in
different
tissues
could
not
be
determined.
The
relatively
low
blood
concentrations
of
BCA
following
oral
dosing,

compared
to
the
blood
concentrations
following
IV
dosing,
suggest
that
the
liver
may
be
an
important
site
for
first­
pass
metabolism.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
14
Draft,
do
not
cite
or
quote
Shultz
et
al.
(
1999)
suggested
that
the
similarities
between
the
toxicokinetics
of
BCA
and
dichloroacetic
acid
(
DCA),
coupled
with
limited
in
vitro
metabolism
data
(
Schultz,
et
al.,
1998;

Tong
et
al.,
1998a),
support
the
hypothesis
that
metabolism
of
BCA
is
similar
to
that
of
DCA.

However,
while
the
metabolic
pathways
for
BCA
and
DCA
might
be
similar,
the
rate
of
metabolism
seems
to
be
greater
for
BCA
than
DCA
based
on
area­
under­
the­
curve
plasma
data.

In
a
review
of
DCA
metabolism,
Stacpoole
et
al.
(
1998)
described
several
potential
mechanisms
for
the
dehalogenation
of
DCA,
but
evidence
for
any
single
pathway
was
limited.

Recent
evidence
suggests
that
DCA
is
metabolized
to
glyoxylic
acid
through
a
glutathione­

Stransferase
dependent
mechanism
by
a
novel
GST
isozyme,
GST­
Zeta
(
Tong
et
al.,
1998b);
in
vitro
studies
have
demonstrated
that
this
enzyme
can
also
catalyze
the
metabolism
of
BCA
to
glyoxylic
acid
(
Tong
et
al.,
1998a).
In
comparative
studies
using
DCA,
Tong
et
al.
(
1998a)
found
that
GST­
Zeta
is
expressed
in
mouse,
rat
and
human­
liver
cytosol.
For
DCA,
the
relative
biotransformation
rates
were
mouse>
rat>
human.
Although
glutathione
was
a
required
cofactor,
it
was
not
consumed
or
oxidized
during
the
conversion
of
DCA
to
glyoxylate.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
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HECD
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15
Draft,
do
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cite
or
quote
Br
Br
O
H
Br
Br
O
OH
OH
O
OH
O
Br
O
O
H
OH
O
O
OH
OH
CO
2
OH
HAOX
Glycolate
GST­
Zeta
OGC
H
H
O
OH
N
H
H
Glycine
OH
H
H
H
O
LDH
Dibromoacetic
acid
Glyoxylic
acid
Oxalic
acid
C
AGT
Figure
III­
1.
Proposed
Metabolism
of
DBAa
AGT
=
alanine:
glyoxylate
aminotransferase
GR
=
glyoxylate
reductase
GSTZ
=
glutathione­
S­
transferase­
Zeta
HAOX
=
glycolate
oxidase
(
2­
hydroxyacid
oxidase)
LDH
=
lactate
dehydrogenase
OGC
=
2­
oxoglutarate:
gloxylate
carboligase
a.
Adapted
largely
from
Kennedy
et
al.
(
1993),
Stacpoole
et
al.
(
1998)
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
16
Draft,
do
not
cite
or
quote
In
addition
to
glyoxylate,
DCA
metabolites
detected
in
vivo
included
monochloroacetic
acid
and
a
series
of
downstream
metabolites
of
glyoxylate
(
Stacpoole
et
al.,
1998).
Thus,
the
formation
of
glyoxylate
metabolites
may
be
of
toxicologic
importance
in
BCA
metabolism
and
is
described
here
briefly.
Glyoxylate
may
be
metabolized
through
competing
pathways
to
form
glycine,
glycolate,
oxalate,
and
CO
2
(
Stacpoole
et
al.,
1998)..
Glyoxylate
can
also
be
converted
to
glycolate
by
glyoxylate
reductase
(
Michal,
1999).
Glycine
may
be
formed
through
the
activity
of
glycolate
aminotransferase
(
formally
known
as
alanine:
glyoxylate
aminotransferase).
Glycine,

which
can
be
incorporated
into
proteins,
used
in
the
synthesis
of
serine,
or
degraded
releasing
carbon
dioxide
(
Michal,
1999).
Conversion
to
oxalate
may
occur
via
a
(
S)­
2­
hydroxyacid
dehydrogenase
such
as
lactate
dehydrogenase
Few
data
are
available
regarding
the
kinetics
of
brominated
acetic
acid
metabolism.
In
the
single
toxicokinetics
study
identified
in
the
literature
that
examined
absorption,
distribution,

metabolism,
and
excretion
(
Schultz
et
al.,
1999),
only
one
dose
of
BCA
was
used
and
analysis
of
metabolic
saturation
could
not
be
conducted.
Gonzalez­
Leon
et
al.
(
1999,
published
abstract)

used
microsomes
to
study
the
effect
of
BCA
pre­
treatment
on
metabolic
inhibition
following
administration
of
subsequent
BCA
doses.
Microsomal
fractions
were
prepared
from
the
livers
of
male
F344
rats
given
2000
mg/
L
BCA
in
drinking
water
for
2
weeks
and
in
vitro
metabolism
was
assessed.
Pretreatment
reduced
the
V
max
by
50%
to
75%,
while
the
K
m
remained
unchanged,

indicating
possible
noncompetitive
inhibition
of
metabolism.
However,
in
the
absence
of
additional
detail
on
laboratory
and
analytic
methods,
it
is
not
clear
whether
metabolism
actually
occurred
in
the
microsomes
or
whether
the
microsomal
samples
were
contaminated
with
cytosol,
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
17
Draft,
do
not
cite
or
quote
metabolism
occurred
in
the
cytosol.
Thus,
an
alternative
interpretation
of
the
data
is
that
the
observed
metabolism
was
due
to
cytosolic
contamination
of
the
microsomal
fraction,
and
that
pretreatment
with
BCA
inhibited
the
cytosolic
enzyme.
Anderson
et
al.
(
1999)
administered
i.
p.

injections
of
0.3
mmol
BCA
to
male
F344
rats
(
3/
dose)
and
measured
GST­
Zeta
activity
in
liver
cytosol
12
hours
later.
BCA
reduced
GST­
Zeta
activity
to
19%
of
that
in
saline­
treated
controls,

indicating
a
possible
mechanism
for
auto­
inhibition
of
metabolism.
The
toxicologic
implications
of
these
findings
are
not
yet
clear
because
it
is
not
known
whether
the
toxic
moiety
is
the
parent
compound
or
a
metabolite.
However,
this
finding
is
of
interest
because
DCA
is
known
to
inhibit
its
own
metabolism
through
inhibition
in
both
rodents
(
Line
et
al.,
1993)
and
humans
(
Stacpoole
et
al.,
1998),
and
BCA
may
also
exhibit
similar
metabolic
activity.
In
the
case
of
DCA
inhibition
is
the
result
of
DCA
reaction
with
and
subsequent
modification
of
GST
zeta.(
Tong
et
al.,
1998a).

Dibromoacetic
acid
As
for
BCA,
most
of
the
information
on
the
metabolism
of
DBA
is
indirect
or
based
on
analogy
to
chlorinated
acetic
acids.
These
indirect
data
suggest
that
DBA
is
likely
to
be
rapidly
metabolized.
Comparisons
of
DBA
renal
and
blood
clearance
in
the
Schultz
et
al.
(
1999)
study
revealed
that
less
than
1%
of
blood
clearance
of
the
parent
compound
was
accounted
for
by
urinary
excretion,
suggesting
that
metabolism
is
the
major
contributor
to
DBA
removal
from
the
blood.
However,
these
data
were
obtained
following
IV
administration
of
the
test
compound
and,

thus,
the
degree
of
metabolism
in
different
tissues
cannot
be
determined
from
the
data
provided.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
18
Draft,
do
not
cite
or
quote
The
limited
oral
bioavailability
of
DBA
suggests
that
the
liver
may
be
an
important
site
for
firstpass
metabolism.
The
proposed
metabolic
pathway
for
DBA,
based
on
the
data
presented
in
the
following
paragraphs,
is
shown
in
Figure
III­
1.

Shultz
et
al.
(
1999)
proposed
that
similarities
between
the
toxicokinetics
of
DCA
and
those
of
DBA
and
BCA
indicate
that
these
compounds
are
likely
to
share
metabolic
pathways.

This
hypothesis
is
supported
by
several
comparative
in
vitro
metabolic
studies
(
Schultz
et
al.,

1998;
Tong
et
al.,
1998a)
that
demonstrated
that
GST­
Zeta
can
catalyze
the
metabolism
of
DBA,

as
well
as
BCA
and
DCA,
to
glyoxylic
acid
(
Tong
et
al.,
1998a)

As
described
in
the
previous
section
for
BCA,
very
little
data
are
available
regarding
the
kinetics
of
brominated
acetic
acid
metabolism.
In
the
single
toxicokinetics
study
identified
in
the
literature
that
examined
absorption
distribution,
metabolism,
and
excretion
(
Schultz
et
al.,
1999),

only
one
dose
of
DBA
was
used.
Therefore,
no
conclusions
regarding
metabolic
saturation
can
be
made.
In
a
study
similar
to
that
reported
in
the
previous
section
for
BCA,
Gonzalez­
Leon
et
al.

(
1999,
published
abstract)
studied
the
effects
of
DBA
pre­
treatment
on
in
vitro
metabolic
inhibition
in
the
liver
microsomes
of
male
F344
rat
liver
following
in
vivo
administration
of
2000
mg/
L
DBA
in
drinking
water
for
2
weeks.
Similar
to
BCA,
pretreatment
reduced
the
V
max
by
50%

to
75%,
while
the
K
m
remained
unchanged,
suggesting
metabolic
inhibition;
however
it
was
not
clear
from
the
abstract
whether
metabolism
occurred
in
the
microsomes
or
in
the
cytosol.

Anderson
et
al.
(
1999)
administered
i.
p.
injections
of
0.3
mmol
DBA
to
male
F344
rats
(
3/
dose)

and
measured
GST­
Zeta
activity
in
liver­
cytosol
preparations
12
hours
later.
DBA
administration
reduced
GST­
Zeta
activity
to
17%
of
that
of
saline­
treated
controls,
indicating
a
possible
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
19
Draft,
do
not
cite
or
quote
mechanism
for
auto­
inhibition
of
metabolism.
The
toxicologic
implications
of
these
findings
are
not
yet
clear
because
it
is
not
known
whether
the
toxic
moiety
is
the
parent
compound
or
a
metabolite.
However,
it
is
of
interest
that
similar
auto­
inhibition
was
seen
for
DCA
metabolism
(
Tong
et
al.,
1998a).

D.
Excretion
Monobromoacetic
acid
No
studies
on
the
excretion
of
MBA
following
exposure
by
any
route
were
identified.

Bromochloroacetic
acid
Schultz
et
al.
(
1999)
measured
parent
compound
concentrations
in
the
blood,
urine,
and
feces
after
oral
or
i.
v.
dosing
of
male
F344
rats
with
500
µ
mol/
kg
(
87
mg/
kg)
BCA.
Blood
measurements
were
taken
at
0,
5,
10,
20,
30,
60
and
90
minutes,
and
at
3,
4,
6,
8
and
12
(
i.
v.

only)
hours
post­
dosing;
urine
and
feces
were
collected
24
hours
following
dosing.
BCA
was
rapidly
cleared
from
the
blood,
with
apparent
bi­
exponential
elimination
following
i.
v.

administration.
There
was
an
initial,
rapid
decline
in
blood
concentration,
corresponding
to
a
short
distributive
phase,
followed
by
a
long
linear
decline.
In
the
concentration­
time
profile,
peculiarities
were
noted
in
the
profile
which
suggested
that
physiological
mechanisms
or
processes
were
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
1
Personal
communication,
I.
Schultz,
Battelle
Laboratories,
Washington
EPA/
OW/
OST/
HECD
III­
20
Draft,
do
not
cite
or
quote
involved
other
than
multiple
distribution
phases
(
e.
g.,
2­
3
compartments
or
distribution
phases).
1
As
a
result,
the
authors
were
doubtful
that
the
unique
appearance
of
the
profiles
was
due
to
a
prolonged
distribution
phase(
s),
and
chose
to
analyze
the
data
using
simple
non­
compartmental
methods,
which
require
fewer
assumptions
than
compartmental
models
with
regard
to
distribution
within
the
animal.
Therefore,
they
provided
two
estimates
of
half­
life
(
t
1/
2):
one
relying
on
the
initial
decline
in
the
profiles
(
0­
4
hours)
and
another
using
the
full
or
complete
profile.
Truncating
the
concentration­
time
profiles
had
no
significant
effect
on
the
AUC;
however,
the
elimination
half­
life
was
markedly
altered
by
more
than
five­
fold;
t
½
was
3.93
hours
for
the
complete
profile
versus
0.74
hours
for
the
truncated
profile.
After
oral
administration,
blood
levels
reached
a
maximum
at
1.5
hours
following
dosing
and
declined
rapidly
during
the
next
6
hours;
t
max
was
1.5
hours.

Removal
of
parent
compound
from
the
blood
appeared
to
be
rapid,
due
mainly
to
biotransformation
(
Schultz
et
al.,
1999).
The
urine
and
feces
were
minimal
contributors
to
overall
blood
clearance.
Urinary
clearance
of
the
parent
compound
accounted
for
2%
of
the
total
clearance
and
feces
contained
negligible
amounts
of
BCA.
The
study
authors
did
not
measure
either
putative
metabolites
or
expired
CO
2.
Therefore,
it
was
not
possible
to
determine
the
contribution
of
each
route
of
excretion
to
the
total
administered
dose
of
the
parent
compound.

No
data
on
the
excretion
of
BCA
following
exposure
by
the
inhalation
or
dermal
routes
were
identified.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
21
Draft,
do
not
cite
or
quote
Dibromoacetic
acid
Schultz
et
al.
(
1999)
also
measured
DBA
concentrations
in
the
blood,
urine,
and
feces
24
hours
after
oral
or
i.
v.
dosing
of
male
F344
rats
with
500
µ
mol/
kg
DBA
(
109
mg/
kg).
Blood
measurements
were
taken
at
0,
5,
10,
20,
30,
60
and
90
minutes,
and
at
3,
4,
6
and
8
hours
postdosing
urine
and
feces
were
collected
24
hours
following
dosing.
DBA
was
rapidly
cleared
from
the
blood,
with
apparent
bi­
exponential
elimination
following
i.
v.
administration.
After
oral
administration,
blood
levels
reached
a
maximum
about
one
hour
following
dosing
and
declined
rapidly
during
the
next
six
hours.
Similar
to
BCA,
features
were
noted
in
the
DBA
concentrationtime
profile
which
suggested
physiological
mechanisms
or
processes
were
involved
other
than
multiple
distribution
phases
(
e.
g.,
2­
3
compartments
or
distribution
phases),
and
the
study
authors
chose
to
analyze
the
data
using
simple
non­
compartmental
methods,
which
require
fewer
assumptions
than
compartmental
models
with
regard
to
distribution
within
the
animal.
As
with
BCA,
they
provided
two
estimates
of
t
1/
2:
one
relying
on
the
initial
decline
in
the
profiles
(
0­
4
hours)
and
another
using
the
full
or
complete
profile.
Unlike
BCA,
DBA
elimination
half­
lives
were
similar
for
both
the
complete
and
truncated
profiles
(
0.72
versus
0.52
hours).
The
urine
and
feces
were
minimal
contributors
to
overall
blood
clearance.
Urinary
clearance
of
DBA
accounted
for
only
a
small
fraction
of
the
total
clearance
and
was
less
than
1%
of
total
clearance;
negligible
amounts
of
DBA
were
found
in
the
feces.
The
study
authors
did
not
measure
either
putative
metabolites
or
expired
CO
2.
Therefore,
it
was
not
possible
to
determine
the
contribution
of
each
route
of
excretion
to
the
total
administered
dose
of
the
parent
compound.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
22
Draft,
do
not
cite
or
quote
No
data
on
excretion
of
DBA
following
exposure
by
the
inhalation
or
dermal
routes
were
identified.

E.
Bioaccumulation
and
Retention
Monobromoacetic
acid
No
studies
on
the
bioaccumulation
and
retention
of
MBA
following
exposure
by
any
route
were
identified.

Bromochloroacetic
acid
The
available
data
on
the
potential
for
bioaccumulation
or
retention
of
BCA
are
very
limited.
Schultz
et
al.
(
1999)
demonstrated
that
a
single
dose
of
BCA
is
rapidly
eliminated
from
the
blood.
Following
intravenous
dosing,
BCA
did
not
appear
to
bind
significantly
to
plasma
proteins
or
accumulate
in
blood
cells.
The
unbound
fraction
in
plasma
and
the
plasma
to
blood
concentration
approached
unity,
and
the
apparent
volume
of
distribution
was
similar
to
the
total
body
water
volume
for
rats
(
Reinoso
et
al.,
1997),
suggesting
that
BCA
distributed
uniformly
outside
the
vascular
system
and
was
unlikely
to
accumulate
significantly
in
peripheral
tissues.

However,
only
a
single
dose
was
administered
and
no
measurements
were
collected
in
tissues
other
than
blood.
Therefore,
the
extent
of
bioaccumulation
or
retention
cannot
be
determined.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
23
Draft,
do
not
cite
or
quote
BCA
at
physiologic
pH
is
not
lipophilic
(
Schultz
et
al.,
1999),
suggesting
little
affinity
for
accumulation
in
adipose
tissue.

No
studies
were
identified
on
the
bioaccumulation
and
retention
of
BCA
following
exposure
by
the
inhalation
or
dermal
routes.

Dibromoacetic
acid
Similar
to
BCA,
DBA
was
rapidly
eliminated
from
the
blood
in
the
toxicokinetic
study
by
Schultz
et
al.
(
1999),
and
did
not
bind
significantly
to
plasma
proteins
or
accumulate
in
blood
cells.
Both
the
unbound
fraction
in
plasma
and
the
plasma:
blood
concentrations
were
close
to
unity
and
the
apparent
volume
of
distribution
was
similar
to
the
total
body­
water
volume
for
rats
(
Reinoso
et
al.,
1997),
suggesting
uniform
distribution
outside
the
vascular
system
and
little
affinity
for
accumulation
in
peripheral
tissues.
However,
as
with
BCA,
only
a
single
dose
was
administered
and
no
measurements
were
collected
in
tissues
other
than
blood.
Rapid
blood
clearance
of
DBA
was
also
suggested
by
the
absence
of
detectable
blood
levels
in
an
NTP
(
1999)

immunotoxicity
study.
Kennedy
et
al.
(
1993)
reported
on
the
tissue
retention
of
radiolabel
following
oral
dosing
of
rats
with
[
14
C]
1,1,2,2­
tetrabromoethane,
a
compound
whose
major
urinary
metabolite
is
DBA.
After
96
hours,
the
percent
of
administered
dose
retained
in
the
body
ranged
from
14%
to
22%
and
was
not
dose­
dependent.
The
largest
percentage
of
dose
was
in
the
carcass,
followed
by
the
liver,
blood,
and
gastrointestinal
tract,
with
lesser
amounts
found
in
the
kidney
and
fat.
These
data
suggest
that
1,1,2,2­
tetrabromoethane
and/
or
one
(
or
more)
of
its
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
24
Draft,
do
not
cite
or
quote
metabolites
were
widely
distributed;
however,
it
is
not
known
whether
the
observed
tissuedistribution
pattern
would
be
similar
following
direct
oral
dosing
with
DBA.
Further,
the
results
of
total
[
14
C]
distribution
did
not
identify
whether
the
tissue
distribution
represented
the
parent
compound
or
its
metabolites,
and
the
specific
metabolites
were
not
identified.

DBA
at
physiologic
pH
is
not
lipophilic
(
Schultz
et
al.,
1999),
suggesting
little
propensity
for
retention
or
accumulation
in
adipose
tissue.
However,
Klinefelter
et
al.
(
2000,
abstract)

reported
the
presence
of
DBA
in
the
milk
of
Sprague­
Dawley
females,
following
high­
dose
exposure
during
pregnancy
and
lactation,
at
concentrations
greater
than
those
measured
in
females'
blood
serum
(
Klinefelter
et
al.,
2000,
abstract),
suggesting
that
retention
or
accumulation
may
be
possible
under
certain
physiologic
conditions.
A
full
report
of
this
study
has
not
been
published
and
thus
these
findings
cannot
be
comprehensively
evaluated.
In
contrast,
Christian
et
al.
(
2001)
did
not
detect
levels
of
DBA
in
the
milk
of
lactating
Sprague­
Dawley
rats
at
drinking
water
concentrations
up
to
1000
ppm
(
estimation
of
daily
doses
to
lactating
dams
was
confounded
by
concomitant
pup
water
consumption).
DBA
was
reliably
detected
on
GD
21
in
placental
tissue
at
1000
ppm
(
81.6
mg/
kg/
day),
and
in
fetal
plasma
at
500
and
1000
ppm
(
maternal
doses
of
48.5
and
81.6
mg/
kg/
day,
respectively),
at
concentrations
which
were
generally
similar
to
those
measured
in
the
plasma
of
pregnant
females
on
GD
20.
Higher
concentrations
of
DBA
in
the
plasma
of
male
and
female
rats
noted
on
SD
14
as
compared
with
SD
1
were
attributed
by
the
authors
to
variability
and
not
to
accumulation.
Placental
DBA
levels
on
GD
21
were
lower
than
those
observed
in
maternal
serum
plasma
on
GD
20
except
at
1000
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
25
Draft,
do
not
cite
or
quote
ppm.
Christian
et
al.
(
2001)
concluded
that
although
DBA
freely
crossed
the
placenta
and
distributed
to
the
fetus
during
gestation,
it
did
not
appear
to
bioaccumulate.

No
studies
were
identified
on
the
bioaccumulation
and
retention
of
DBA
following
exposure
by
the
inhalation
or
dermal
routes.

F.
Summary
No
toxicokinetic
studies
of
MBA
have
been
identified
in
the
literature.
Brominated
acetic
acids
appear
to
be
rapidly
absorbed
from
the
gastrointestinal
tract
(
Schultz
et
al.,
1999).
Key
data
from
this
study
are
summarized
in
Table
III­
1.
Following
single­
dose
intravenous
or
oral­
gavage
exposure,
both
BCA
and
DBA
were
rapidly
cleared
from
blood
and
had
short
plasma
elimination
half­
lives
(
Schultz
et
al.,
1999).
However,
the
extent
of
tissue
distribution
is
not
known
because
tissue
distribution
studies
of
these
compounds
have
not
been
conducted.
Neither
BCA
nor
DBA
are
lipophilic
at
physiologic
pH
(
Schultz
et
al.,
1999),
suggesting
a
low
propensity
to
accumulate
in
fat.
Following
repeated
exposure
in
drinking
water,
DBA
was
detected
in
the
blood
serum
and
milk
of
lactating
female
Sprague­
Dawley
rats
exposed
from
gestation
day
15
to
postnatal
day
20
(
Klinefelter,
2000,
abstract),
but
was
not
detected
in
the
blood
plasma
of
nonlactating
female
B6C3F1
mice
after
28
days
of
exposure
(
NTP,
1999).
The
discrepancy
in
these
findings
might
be
due
to
sampling
differences,
species
differences,
and/
or
differences
in
physiologic
status.

The
metabolism
of
BCA
and
DBA
has
also
not
been
thoroughly
investigated.
Limited
in
vitro
data
(
Schultz
et
al.,
1998;
Tong
et
al.,
1998a,
1998b)
and
the
single
comparative
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
III­
26
Draft,
do
not
cite
or
quote
toxicokinetics
study
by
Schultz
et
al.
(
1999)
suggest
that
both
BCA
and
DBA
are
metabolized
in
a
manner
similar
to
DCA.
Potential
pathways
of
brominated
acetic­
acid
metabolism
to
glyoxylic
acid
have
been
proposed
based
on
analogy
to
chlorinated
acetic
acids
(
reviewed
in
Stacpoole
et
al.,
1998);
these
pathways
are
mediated
through
a
recently­
identified
class
of
GST
isoenzymes,

GST­
Zeta.
It
is
not
clear
whether
the
effective
toxicologic
moiety
is
the
parent
compound
or
an
active
metabolite.
Overall,
the
data
are
consistent
with
DBA
and
DCA
being
rapidly
excreted
and
having
little
propensity
for
bioaccumulation.
Lactational
exposure
may
be
a
route
of
concern
because
of
the
presence
of
DBA
in
the
milk
of
lactating
Sprague­
Dawley
females
at
concentrations
greater
than
those
measured
in
the
females'
blood
serum
(
Klinefelter
et
al.,
2000,

abstract).
In
contrast,
Christian
et
al.
(
2001)
did
not
detect
DBA
in
the
milk
samples
of
lactating
Sprague­
Dawley
rats
exposed
to
high
DBA
drinking
water
concentrations,
although
DBA
was
measurable
in
fetal
plasma
on
gestation
day
21.
Christian
et
al.
(
2001)
observed
that,
although
DBA
freely
crossed
the
placenta
in
pregnant
Sprague­
Dawley
rats,
the
attained
maternal
and
fetal
plasma
levels
were
associated
with
the
amount
and
timing
of
water
consumption
and
did
not
appear
to
accumulate.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
1
Draft,
do
not
cite
or
quote
Chapter
IV.
Human
Exposure
A.
Drinking
Water
Exposure
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.

MBA,
BCA,
and
DBA
have
been
identified
as
drinking­
water
disinfection
byproducts
under
the
Information
Collection
Rule
(
U.
S.
EPA,
1994)
and
are
being
assessed
for
regulatory
consideration
in
the
Stage
2
Disinfectants/
Disinfection
Byproducts
Rule
to
be
promulgated.

Therefore,
this
section
will
examine
the
occurrence
of
these
compounds
in
drinking
water.

A.
1
National
Occurrence
Data
for
MBA,
BCA,
and
DBA
This
section
presents
the
data
collected
from
the
Information
Collection
Rule
(
ICR)

databases,
which
provide
data
from
surface­
and
ground­
water
systems
serving
at
least
100,000
persons.
This
data
base
includes
information
gathered
for
18
months
from
July
1997
to
December
1998.

Section
A.
1.1
describes
the
ICR
data
set
and
analysis
techniques
used
to
present
the
data
for
the
plants
that
submitted
data
under
the
ICR.
The
data
in
Sections
A.
1
and
A.
2
were
taken
from
the
online
version
of
the
ICR
database
(
U.
S.
EPA,
2000a),
and
the
explanation
of
the
methods
used
was
taken
from
the
Draft
EPA
Document
on
Stage
2
Occurrence
and
Exposure
Assessment
for
Disinfectants
and
Disinfection
Byproducts
(
D/
DBPs)
in
Public
Drinking
Water
(
U.
S.
EPA,
2000b).
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
2
Draft,
do
not
cite
or
quote
A.
1.1
ICR
Plants
The
ICR
generated
plant­
level
sets
of
data
that
link
water
quality
and
treatment
from
source
to
tap,
and
aid
in
understanding
the
seasonal
variability
in
these
relationships.
The
database
contains
information
from
18
monthly
or
6
quarterly
samples
from
7/
97
to
12/
98
from
approximately
300
large
systems
covering
approximately
500
plants.
These
samples
were
tested
for
influent
and
finished
water­
quality
parameters
(
e.
g.,
TOC,
temperature,
pH,
alkalinity),
DBP
levels,
and
disinfectant
residuals.
Samples
were
collected
at
several
locations
throughout
the
distribution
system
to
cover
the
entire
range
of
residence
times
during
which
DBPs
can
form
in
the
finished
water.
Over
the
18­
month
period,
approximately
1470
samples
were
taken
from
305
plants
with
surface
water
as
their
source,
and
approximately
580
samples
were
taken
from
123
plants
with
groundwater
as
their
source.
For
more
detailed
information,
such
as
sampling
locations
and
frequencies,
refer
to
the
ICR
Data
Analysis
Plan
(
U.
S.
EPA,
2000c).

A.
1.2
Quarterly
Distribution
System
Average
and
Highest
Value
for
MBA,
BCA,
and
DBA
This
section
describes
the
data­
analysis
techniques
employed
for
the
analysis
of
observed
data
for
water­
quality
parameters,
and
for
MBA,
BCA,
and
DBA
concentrations.
All
data
are
categorized
according
to
the
types
of
source
water
­
surface
or
ground.
Plants
having
both
surface­
and
ground­
water
sources
(
mixed)
or
that
purchase
water
are
included
in
the
surface­
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
3
Draft,
do
not
cite
or
quote
water
category.
Quarterly
Distribution
System
Average
and
Highest
Value
for
the
brominated
acetic
acids
are
presented
in
Table
IV­
1.

The
quarterly
distribution­
system
average
is
an
average
of
the
following
four
distinct
locations
in
the
distribution
system.


Distribution
System
Equivalent
(
DSE)
location;


Average
1
(
AVG
1)
and
Average
2
(
AVG
2)
locations:
Two
sample
points
in
the
distribution
system
representing
the
approximate
average
residence
time
as
designated
by
the
water
system;
and

Distribution
System
Maximum:
Sample
point
in
the
distribution
system
having
the
highest
residence
time
(
or
approaching
the
longest
time)
as
designated
by
the
water
system
The
quarterly
distribution­
system
highest
value
is
the
highest
of
the
four
distributionsystem
samples
collected
by
a
plant
in
a
given
quarter.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
4
Draft,
do
not
cite
or
quote
Table
IV­
1.
Bromoacetic
Acids
Quarterly
Distribution
System
Average
and
Highest
Value
Source
Quarterly
Dist.
Sys.
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
MBA
SW
Average
305
1470
83.06
0.25
0.00
0.98
ND
26.00
0.00
0.96
High
305
1470
83.06
0.41
0.00
1.84
ND
58.10
0.00
1.50
GW
Average
123
581
86.57
0.16
0.00
0.54
ND
6.68
0.00
0.38
High
123
581
86.57
0.27
0.00
0.92
ND
12.00
0.00
1.30
BCA
SW
Average
305
1474
9.70
3.61
2.88
3.08
ND
24.18
0.25
7.70
High
305
1474
9.70
4.45
3.50
3.86
ND
41.90
1.00
9.00
GW
Average
123
584
48.63
1.47
0.28
2.15
ND
11.28
0.00
4.50
High
123
584
48.63
2.18
1.10
3.40
ND
41.00
0.00
6.40
DBA
SW
Average
305
1484
60.58
1.09
0.00
2.08
ND
14.25
0.00
3.68
High
305
1484
60.58
1.39
0.00
2.52
ND
19.00
0.00
4.30
GW
Average
123
584
56.16
0.82
0.00
1.48
ND
13.00
0.00
2.58
High
123
584
56.16
1.20
0.00
1.88
ND
16.00
0.00
3.70
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Quarterly
Dist.
Sys:
Quarterly
Distribution
System
Average
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
ND:
Nondetected
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
5
Draft,
do
not
cite
or
quote
The
mean
concentrations
of
MBA
(
averaged
across
the
four
sampling
locations)
were
0.16
and
0.25

g/
L
in
groundwater
and
surface
water,
respectively.
The
mean
concentrations
of
BCA
(
averaged
across
the
four
sampling
locations)
were
1.47
and
3.61

g/
L
in
groundwater
and
surface
water,
respectively.
The
mean
concentrations
of
DBA
(
averaged
across
the
four
sampling
locations)
were
0.82
and
1.09

g/
L
in
groundwater
and
surface
water,
respectively.
Examination
of
the
data
using
the
Student's
t­
test
indicates
that
the
mean
concentrations
of
MBA,
BCA,
and
DBA
in
surface
water
was
significantly
higher
that
the
mean
concentrations
of
these
chemicals
in
ground
water.
The
mean
concentrations
of
BCA
were
statistically
significantly
higher
(
p
=
0.05)

than
the
mean
concentrations
of
DBA,
which
were
significantly
higher
(
p
=
0.05),
that
the
mean
MBA
concentrations
in
both
surface­
and
ground­
water.
The
lowest
mean
concentrations
are
associated
with
the
highest
percentage
of
nondetects,
which
are
treated
as
0
in
the
calculation
of
the
mean,
median,
standard
deviation,
and
p10
values
(
U.
S.
EPA,
2000a).

A.
2
Factors
Affecting
the
Relative
Concentrations
of
MBA,
BCA,
and
DBA
in
Drinking
Water
Sections
A.
2.1
­
A.
2.4
contain
investigational
information
and
ICR
data
on
the
effects
of
disinfection
chemicals,
influent
bromide
concentration,
influent
total
organic
carbon
(
TOC)

concentration,
and
seasonal
shifts,
respectively
in
MBA,
BCA,
and
DBA
concentrations.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
6
Draft,
do
not
cite
or
quote
A.
2.1
Disinfection
Treatment
Chlorination
has
been
the
predominant
water­
disinfection
method
in
the
United
States.

However,
water
utilities
are
considering
a
shift
to
alternative
disinfectants.
Therefore,
there
is
a
need
to
understand
the
occurrence
of
DBPs
in
drinking
water
and
the
factors
that
may
influence
their
formation.
Several
published
studies
(
Boorman
et
al.,
1999;
Richardson,
1998;
Lykins
et
al.,

1994;
Jacangelo
et
al.,
1989)
reported
on
the
formation
of
brominated
acetic
acids
and
other
DBPs
under
different
disinfection
conditions.

In
a
review
on
drinking­
water
disinfection
byproducts,
Boorman
et
al.
(
1999)
compared
the
concentrations
of
different
drinking­
water
disinfection
byproducts,
including
MBA,
BCA,
and
DBA,
formed
by
chlorination,
ozonation,
chlorine
dioxide,
and
chloramination.
Most
of
the
data
were
available
for
surface­
water
systems
that
used
chlorination.
For
the
systems
using
chlorination,
BCA,
with
a
median
and
a
maximum
concentration
of
3.2
and
49
µ
g/
L,
respectively,

was
present
at
the
highest
concentrations.
The
median
values
of
both
MBA
and
DBA
in
chlorinated
water
were
less
than
0.5
µ
g/
L,
with
maximum
values
of
1.7
and
7.4
µ
g/
L,

respectively.
The
principal
products
formed
by
chloramination
were
similar
to
those
formed
by
chlorination;
additional
information
was
not
provided.
Ozonation
of
water
containing
bromide
may
produce
DBPs
such
as
bromoform,
dibromoacetic
acid,
cyanogen
bromide,
and
bromate.
The
total
concentration
of
brominated
acetic
acids
formed
by
the
use
of
ozonation
ranged
from
1
to
50
µ
g/
L;
concentrations
for
individual
compounds
were
not
provided.
Chlorine
dioxide
formed
oxidation
by­
products
similar
to
those
formed
by
ozonation;
additional
details
were
not
provided.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
7
Draft,
do
not
cite
or
quote
Richardson
(
1998)
compared
the
relative
concentrations
of
DBPs
in
drinking
water
using
different
treatment
methods,
and
found
that
chlorination
produced
the
highest
concentration
of
DBPs,
including
MBA,
BCA,
and
DBA.
Chlorine
dioxide
and
chloramine,
when
compared
to
chlorine,
produced
fewer
and
lower
concentrations
of
DBPs.
MBA,
BCA,
and
DBA
were
not
produced
by
chlorine
dioxide
in
measurable
quantities.
Compared
to
chlorine
treatment,

chloramine
produced
3%
to
20%
lower
levels
of
by­
products,
including
haloacetic
acids.

Ozonation
produced
insignificant
levels
of
trihalomethanes
(
THMs).
However,
when
elevated
levels
of
bromide
ion
were
in
the
raw
water,
MBA
and
DBA
were
detected
following
ozonation.

When
ozone
was
the
primary
disinfectant
followed
by
chloramine,
the
levels
of
most
DBPs,

including
haloacetic
acids,
were
lower
than
when
chloramine
was
used
solely.
However,
there
was
an
observed
shift
to
more
brominated
species
of
THMs
and
haloacetic
acids
when
ozone
was
followed
by
chlorine,
than
when
chorine
was
used
solely.

Lykins
et
al.
(
1994)
investigated
the
formation
of
halogenated
DBPs
in
the
waterdistribution
system,
by
predisinfecting
and
postdisinfecting
the
water
with
either
chlorine
or
chloramine
and
holding
the
water
for
five
days.
They
found
that
the
use
of
chlorine
produced
the
highest
concentration
of
halogenated
DBPs
and
that,
in
general,
the
concentrations
could
be
reduced
by
adding
ozone
as
a
predisinfectant
with
postchlorination.
Lykins
et
al.
(
1994)
also
found
that
the
highest
average
concentrations
of
MBA
(
1.2

g/
L)
and
BCA
(
18

g/
L)
were
formed
when
chlorine
was
the
sole
treatment
method.
The
average
DBA
concentration
was
0.6

g/
L
when
chlorine
was
the
sole
treatment
method.
The
next
highest
concentrations
of
MBA
(
1.0

g/
L)
and
BCA
(
14

g/
L)
were
observed
with
ozone
treatment
followed
by
chlorine.
In
contrast,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
8
Draft,
do
not
cite
or
quote
the
average
DBA
concentration
(
1.0

g/
L)
was
slightly
higher
with
ozone
treatment
followed
by
chlorine
than
when
chlorine
was
the
sole
treatment
method.
Chloramine
treatment,
and
ozone
treatment
followed
by
chloramine,
resulted
in
the
lowest
concentrations
of
MBA
and
DBA,
with
both
treatment
methods
resulting
in
0.1

g/
L
MBA
and
0.1

g/
L
DBA.
BCA
concentrations
were
lower
with
ozone
treatment
followed
by
chloramine
(
1.0

g/
L
BCA)
than
when
chloramine
was
the
sole
treatment
method
(
1.9

g/
L).

Jacangelo
et
al.
(
1989)
examined
the
impact
of
ozonation
on
the
formation
and
control
of
DBPs
in
drinking
water
at
four
utilities.
Treatment
modifications
were
made
on
the
process
train
at
each
full
or
pilot­
scale
plant
to
incorporate
ozone
in
the
treatment
process.
For
two
of
the
utilities
in
the
Jacangelo
et
al.
(
1989)
study,
only
total
haloacetic
acids
(
HAAs)
were
measured,

and
no
measurements
were
made
of
individual
HAAs.
The
disinfection
schemes
that
employed
ozonation
followed
by
chloramines
as
a
disinfectant
resulted
in
large
decreases
in
HAAs
relative
to
chlorination.
However,
the
sample
size
did
not
allow
for
statistical
analysis
of
the
data
(
Jacangelo
et
al.,
1989).
For
two
other
utilities
that
measured
individual
HAAs,
preozonation
followed
by
chlorination
decreased
the
total
HAAs
by
14
to
50%,
when
compared
with
chlorination
only.
The
concentration
of
MBA
was
essentially
the
same
with
and
without
preozonation,
while
DBA
increased
with
ozonation.
BCA
was
not
measured
in
this
study.
The
authors
suggested
that
ozone
reacts
with
bromide
ions
in
the
source
water,
resulting
in
the
formation
of
hypobromous
acid.
Reaction
of
hypobromous
acid
and
natural
organic
matter
can
produce
brominated
HAAs.
When
preozonation
and
postchlorination
are
practiced,
competition
exists
between
hypochlorous
acid
and
hypobromous
acid
for
organic
matter,
leading
to
varying
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
9
Draft,
do
not
cite
or
quote
concentrations
of
chlorinated
and
brominated
HAAs
(
Jacandgelo
et
al.,
1989).
In
addition,

Jacangelo
et
al.
(
1989)
noted
that
the
concentrations
of
brominated
acetic
acids
increased
with
increasing
bromide
concentrations
in
the
source
water.

Miltner
et
al.
(
199)
studied
DBP
formation
and
control
in
three
surface
water
pilot
plants
employing
three
different
disinfectant
methods
(
chlorine,
ozone
followed
by
chlorine,
and
ozone
followed
by
chloramine).
In
an
examination
of
the
data
using
the
Student's
t­
test,
the
authors
found
that
the
amount
of
BCA
measured
in
finished
water
and
in
simulated
distribution
waters
was
lower
(
p
=
0.05)
when
ozonation
was
combined
with
chlorination
or
with
cloramination
than
when
chlorination
was
used
a
lone.
However,
ozonation
had
no
effect
(
p
=
0.05)
on
the
formation
of
MBA,
and
the
formation
of
DBA
was
higher
(
p
=
0.05)
when
ozonation
was
followed
by
chlorination
than
when
chlorination
alone
was
used.

A.
2.1.1
Disinfection
Treatment
in
ICR
Database
Data
on
the
concentrations
of
MBA,
BCA,
and
DBA
were
gathered
from
plants
using
several
disinfection
treatments.
Those
chemical­
disinfection
treatments
most
commonly
used
(
used
by
10%
or
more
of
the
plants
evaluated),
along
with
the
ozonation
treatments,
are
presented
in
Tables
IV­
2,
IV­
3,
and
IV­
4
for
MBA,
BCA,
and
DBA,
respectively.

Examination
of
the
data
using
the
Student's
t­
test
indicates
that,
for
all
chemicaldisinfection
treatments
used
for
surface
water,
the
mean
concentrations
of
BCA
were
statistically
significantly
higher
(
p
=
0.05)
than
the
mean
concentrations
of
DBA
which
were
significantly
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
10
Draft,
do
not
cite
or
quote
higher
(
p
=
0.05)
than
the
mean
concentrations
(
Tables
IV­
2
and
IV­
3).
For
all
chemicaldisinfection
treatments
used
for
ground
water,
the
mean
concentrations
of
BCA
were
significantly
higher
(
p
=
0.05)
than
the
mean
concentrations
of
MBA.
The
mean
concentrations
of
BCA
in
groundwater
were
significantly
higher
(
p
=
0.05)
than
the
mean
DBA
concentrations
only
in
plants
using
ozone
and
choramine.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
11
Draft,
do
not
cite
or
quote
Table
IV­
2.
MBA
by
Disinfection
Method
(
Quarterly
Distribution
System
Average)

Source
Disinfection
Chemicals
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
Cl
2/
Cl
2
180
814
85.63
0.24
0.00
1.14
ND
26.00
0.00
0.75
Cl
2_
CLM/
CLM
66
307
80.78
0.27
0.00
0.72
ND
5.73
0.00
1.10
O
3/
CL
2
7
25
88.00
0.11
0.00
0.33
0.00
1.28
0.00
0.28
O
3/
CLM
10
49
85.71
0.10
0.00
0.30
0.00
1.48
0.00
0.58
GW
/
Cl
2
67
299
87.29
0.18
0.00
0.64
ND
6.68
0.00
0.35
Cl
2/
Cl
2
39
170
85.88
0.15
0.00
0.45
ND
2.53
0.00
0.51
O
3/
CLM
1
6
100.0
0
0.00
0.00
0.00
0.00
0.00
0.00
0.00
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Cl2/
Cl2:
Free
chlorine
in
Water
Treatment
Plant
(
WTP)
and
Distribution
System
(
DS).

Cl2_
CLM/
CLM:
Free
chlorine
followed
by
chloramine
in
WTP
and
chloramine
in
DS.
/
Cl2:
No
disinfectant
in
WTP
and
free
chlorine
in
DS.

O3/
Cl2:
Ozone
in
WTP
and
free
chlorine
in
DS.

O3/
CLM:
Ozone
in
WTP
and
chloramine
in
DS.
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
12
Draft,
do
not
cite
or
quote
Table
IV­
3.
BCA
Acid
by
Disinfection
Method
(
Quarterly
Distribution
System
Average)

Source
Disinfection
Chemicals
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
Cl
2/
Cl
2
180
816
11.40
3.14
2.54
2.77
ND
18.73
0.00
6.83
Cl
2_
CLM/
CLM
66
306
1.63
4.67
4.01
3.28
ND
23.80
1.35
8.83
O
3/
CL
2
7
25
4.00
2.54
2.13
1.77
0.00
5.68
0.33
4.90
O
3/
CLM
10
49
22.45
2.19
2.15
1.70
0.00
6.60
0.00
4.55
GW
/
Cl
2
67
301
65.78
0.87
0.00
1.77
ND
10.70
0.00
2.70
Cl
2/
Cl
2
39
170
45.53
1.27
0.31
1.80
ND
7.88
0.00
3.78
O
3/
CLM
1
6
0.00
2.68
2.55
0.69
2.05
3.88
2.05
3.88
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Cl2/
Cl2:
Free
chlorine
in
Water
Treatment
Plant
(
WTP)
and
Distribution
System
(
DS).

Cl2_
CLM/
CLM:
Free
chlorine
followed
by
chloramine
in
WTP
and
chloramine
in
DS.
/
Cl2:
No
disinfectant
in
WTP
and
free
chlorine
in
DS.

O3/
Cl2:
Ozone
in
WTP
and
free
chlorine
in
DS.

O3/
CLM:
Ozone
in
WTP
and
chloramine
in
DS.
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
13
Draft,
do
not
cite
or
quote
Table
IV­
4.
DBA
by
Disinfection
Method
(
Quarterly
Distribution
System
Average)

Source
Disinfection
Chemicals
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
Cl
2/
Cl
2
180
823
68.53
0.68
0.00
1.55
ND
12.50
0.00
2.20
Cl
2_
CLM/
CLM
66
308
51.62
1.36
0.00
2.11
ND
14.25
0.00
4.83
O
3/
CL
2
7
25
68.00
1.02
0.00
2.27
0.00
8.15
0.00
2.83
O
3/
CLM
10
49
46.94
1.01
0.25
1.50
0.00
7.30
0.00
2.80
GW
/
Cl
2
67
303
67.66
0.68
0.00
1.52
ND
13.00
0.00
2.53
Cl
2/
Cl
2
39
169
46.15
1.01
0.25
1.62
ND
7.25
0.00
3.15
O
3/
CLM
1
6
50.00
0.41
0.14
0.55
0.00
1.25
0.00
1.25
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Cl2/
Cl2:
Free
chlorine
in
Water
Treatment
Plant
(
WTP)
and
Distribution
System
(
DS).

Cl2_
CLM/
CLM:
Free
chlorine
followed
by
chloramine
in
WTP
and
chloramine
in
DS.
/
Cl2:
No
disinfectant
in
WTP
and
free
chlorine
in
Distribution
System.

O3/
Cl2:
Ozone
in
WTP
and
free
chlorine
in
DS.

O3/
CLM:
Ozone
in
WTP
and
chloramine
in
DS.
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
Examination
of
the
ICR
data
using
the
Student's
t­
test
indicates
that
the
mean
concentrations
of
BCA
and
DBA
were
significantly
higher
(
p
=
0.05)
in
surface
water
plants
using
chlorine
followed
by
chloramine
than
in
those
using
free
chlorine
alone.
There
were
no
significant
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
14
Draft,
do
not
cite
or
quote
differences
(
p
=
0.05)
between
the
mean
concentrationof
MBA
in
surface
water
plants
using
chlorine
followed
by
chloramine
and
the
concentration
those
solely
using
chlorine.

In
plants
with
groundwater
as
a
source,
the
mean
concentrtions
of
BCA
and
DBA
in
plants
with
no
disinfectant
in
the
treatment
plant
and
with
free
chlorine
in
the
distribution
system
were
significantly
lower
(
p
=
0.05)
than
the
mean
concentrations
of
the
same
chemicals
in
plants
with
free
chlorine
in
both
the
treatment
plant
and
the
distribution
system.
There
were
no
significant
differences
(
p
=
0.05)
between
the
mean
concentrtions
of
MBA
in
plants
with
no
disinfectant
in
the
treatment
plant
and
with
free
chlorine
in
the
distribution
system
and
those
with
free
chlorine
in
both
the
treatment
plant
and
the
distribution
system.

Only
a
very
limited
number
of
plants
used
ozonation
in
combination
with
either
chlorine
or
chloramine
in
the
distribution
system.
An
examination
of
the
ICR
data
in
Tables
IV­
2,
IV­
3
and
IV­
4
using
the
Student's
t­
test
indicates
that,
with
one
exception,
the
mean
concentrations
of
BCA
were
significantly
lower
in
surface­
water
plants
that
use
ozone
in
the
water­
treatment
plant
and
free
chlorine
or
free
chloramine
in
the
distribution
system
than
in
plants
using
non­
ozonation
disinfection
processes.
This
finding
was
also
presented
by
Lykins
et
al.
(
1994).
However
,
there
were
no
significant
differences
(
p
=
0.05)
between
the
mean
concentrations
of
BCA,
in
surfacewater
plants
when
free
chlorine
was
used
solely,
and
the
BCA
concentrtions
when
both
ozone
and
chlorine
were
used.
There
were
no
significant
differences
(
p
=
0.05)
between
the
concentrations
of
MBA
and
DBA
in
surface­
water
plants
using
common
(
non­
ozonation)
disinfection
processes
and
the
concentrations
of
the
same
chemicals
pin
plants
with
ozone
in
the
water­
treatment
plant
and
free
chlorine
or
free
chloramine
in
the
distribution
system.
In
addition,
there
were
no
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
15
Draft,
do
not
cite
or
quote
significant
differences
(
p
=
0.05)
for
MBA,
BCA,
or
DBA
between
the
two
treatments
using
ozonation
in
treating
surface
water.

There
was
only
1
groundwater
plant
that
used
the
ozone/
chloramine
disinfection
method
and
statistical
analysis
was
not
conducted.

In
summary,
an
analysis
of
the
ICR
data
suggest
that
although
BCA
concentrations
in
surface
water
treated
with
chlorine
are
similar
to
those
treated
with
ozone
and
chlorine,
surface
water
plants
using
ozonation
had
lower
BCA
concentrations
than
those
using
most
common
(

nonozonation
disinfection
processes.
In
addition,
ozonation
appeared
to
have
no
effect
on
the
formation
of
MBA
and
DBA.

A.
2.2
Bromide
Concentration
Pourmoghaddas
et
al.
(
1993)
examined
the
effects
of
source
water
and
treatment
characteristics,
such
as
pH,
reaction
time,
chlorine
dosage,
and
bromide­
ion
concentration,
on
the
formation
of
HAAs.

The
study
quantified
nine
HAA
species
in
the
presence
of
bromide
ion
at
low,
neutral,
and
high
pH
over
time
at
two
chlorine
dosages.
This
study
found
a
shift
in
the
distribution
of
HAAs
from
chlorinated
to
brominated
and
mixed
(
bromochlorinated)
halogenated
species
with
increased
bromide­
ion
concentration.
Chloride­
ion
concentration
had
no
observed
effect
on
the
formation
of
brominated
HAAs.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
16
Draft,
do
not
cite
or
quote
At
the
low
chlorine
dose
of
11.5
mg/
L,
MBA
showed
a
consistent
trend
toward
higher
concentrations
as
reaction
time
and
bromide­
ion
concentrations
increased.
The
only
apparent
effect
of
pH
was
to
increase
the
amount
of
MBA
at
the
highest
bromide
concentration
(
4.5
mg/
L).
The
highest
measured
concentration
of
MBA
was
at
a
pH
of
5
and
a
bromide
concentration
of
4.5
mg/
L
(
Pourmoghaddas
et
al.,
1993).
DBA
concentrations
increased
with
increasing
bromide­
ion
concentration
and
reaction
time.
Changes
in
pH
had
little
influence
on
the
formation
of
DBA
(
Pourmoghaddas
et
al.,
1993).

BCA
was
formed
only
if
the
bromide
ion
was
present.
BCA
concentrations
increased
with
reaction
time
and
were
significantly
lower
at
pH
9.4.
The
highest
observed
concentration
of
BCA
was
at
1.5
mg/
L
bromide­
ion
concentration,
while
BCA
levels
decreased
at
the
highest
bromide
concentration
of
4.5
mg/
L
(
Pourmoghaddas
et
al.,
1993).
It
is
possible
that
the
decreased
BCA
levels
at
high
bromide
concentrations
reflect
the
preferential
formation
of
DBA
over
BCA
under
such
conditions,
but
the
data
provided
are
insufficient
to
test
this
hypothesis.

A.
2.2.1
Bromide
Concentration
in
ICR
Database
Tables
IV­
5,
IV­
6,
and
IV­
7
present
the
formation
of
MBA,
BCA,
and
DBA,
respectively,

as
a
function
of
influent
bromide
concentrations.

Bromide
concentrations
tended
to
be
lower
in
plants
using
surface
water
as
a
source
than
in
those
using
groundwater
as
a
source.
For
example,
114
of
the
305
plants
using
surface
water
as
the
source
(
37%)
had
influent
bromide
levels
below
the
minimal
reporting
limit
(
MRL)
of
20
ppb,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
17
Draft,
do
not
cite
or
quote
while
only
13
of
the
123
plants
using
groundwater
as
the
source
(
11%)
had
influent
bromide
levels
below
the
MRL.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
18
Draft,
do
not
cite
or
quote
Table
IV­
5.
MBA
by
Influent
Bromide
Concentration
(
Quarterly
Distribution
System
Average)

Source
Influent
Bromide
Conc.
(
ppb)
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
<
MRL
(
20)
114
556
90.83
0.14
0.00
0.59
ND
6.63
0.00
0.00
20
­
<
30
41
200
88.50
0.21
0.00
0.72
ND
4.03
0.00
0.38
30
­
<
50
48
221
83.26
0.45
0.00
2.02
ND
26.00
0.00
1.13
50
­
<
100
59
282
77.30
0.24
0.00
0.57
ND
3.95
0.00
1.03

100
39
192
64.06
0.41
0.00
1.12
ND
7.20
0.00
1.90
GW
<
MRL
(
20)
13
66
93.94
0.07
0.00
0.30
ND
1.60
0.00
0.00
20
­
<
30
11
50
98.00
0.01
0.00
0.04
ND
0.30
0.00
0.00
30
­
<
50
26
109
96.33
0.03
0.00
0.18
ND
1.45
0.00
0.00
50
­
<
100
32
148
83.11
0.18
0.00
0.67
ND
6.68
0.00
0.63

100
41
208
78.85
0.27
0.00
0.66
ND
3.65
0.00
1.20
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
MRL:
Minimum
Reporting
Limit
ND:
Nondetect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
19
Draft,
do
not
cite
or
quote
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
20
Draft,
do
not
cite
or
quote
Table
IV­
6.
BCA
by
Influent
Bromide
Concentration
(
Quarterly
Distribution
System
Average)

Source
Influent
Bromide
Conc.
(
ppb)
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
<
MRL
(
20)
114
559
22.00
1.76
1.65
1.79
ND
23.80
0.00
3.43
20
­
<
30
41
197
0.51
3.32
3.03
1.67
ND
12.28
1.63
5.30
30
­
<
50
48
225
3.11
3.88
3.40
2.94
ND
22.50
0.93
7.18
50
­
<
100
59
282
3.90
5.43
4.96
3.26
ND
15.50
1.48
9.68

100
39
192
0.52
6.29
6.01
3.52
ND
24.18
2.60
11.20
GW
<
MRL
(
20)
13
66
78.79
0.35
0.00
1.34
ND
10.25
0.00
0.75
20
­
<
30
11
51
58.52
0.69
0.00
1.19
ND
6.00
0.00
2.25
30
­
<
50
26
108
49.07
1.16
0.25
1.61
ND
7.05
0.00
3.58
50
­
<
100
32
150
48.67
1.22
0.26
1.86
ND
7.88
0.00
4.23

100
41
209
36.36
2.35
1.83
2.60
ND
11.28
0.00
6.25
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
MRL:
Minimum
Reporting
Limit
ND:
Nondetect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
21
Draft,
do
not
cite
or
quote
Table
IV­
7.
DBA
by
Influent
Bromide
Concentration
(
Quarterly
Distribution
System
Average)

Source
Influent
Bromide
Conc.
(
ppb)
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
<
MRL
(
20)
114
561
91.62
0.08
0.00
0.32
ND
3.00
0.00
0.00
20
­
<
30
41
201
67.66
0.29
0.00
0.59
ND
3.48
0.00
1.10
30
­
<
50
48
226
58.41
0.66
0.00
1.10
ND
9.58
0.00
2.03
50
­
<
100
59
284
32.04
1.70
1.33
1.84
ND
7.65
0.00
4.30

100
39
193
6.74
4.37
3.78
3.22
ND
14.25
0.58
8.48
GW
<
MRL
(
20)
13
67
88.06
0.09
0.00
0.31
ND
1.68
0.00
0.28
20
­
<
30
11
50
90.00
0.12
0.00
0.49
ND
2.45
0.00
0.13
30
­
<
50
26
109
69.72
0.27
0.00
0.54
ND
2.15
0.00
1.10
50
­
<
100
32
150
42.00
1.00
0.40
1.27
ND
5.83
0.00
3.10

100
41
208
40.87
1.38
0.60
2.00
ND
13.00
0.00
3.73
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
MRL:
Minimum
Reporting
Limit
ND:
Nondetect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
22
Draft,
do
not
cite
or
quote
A
regression
analysis
of
the
ICR
data
indicates
that,
with
the
exception
of
MBA
in
surface
water,
there
was
a
significant
correlation
(
 
=
0.05)
between
influent
bromide
concentration
and
the
mean
concentrations
of
BCA
and
DBA
in
surface
water
and
groundwater.(
Tables
IV­
5,
IV­
6
and
IV­
7).

Tan
examination
of
th
data
using
the
Student's
t­
test
indicates
that
in
plants
treating
surface
water,
the
mean
concentrations
of
BCA
were
significantly
higher
(
p
=
0.05_
than
the
mean
MBA
and
DBA
concentrtions,
for
a
given
bromide
concentration
range
(
Tables
IV­
5,
IV­
6,

and
IV­
7).
In
addition,
the
mean
concentrations
of
DBA
were
significantly
lower
(
p
=
0.05)
than
the
mean
MBA
concentrations
at
the
lowest
influent
bromide
concentration
(<
minimum
reporting
limit
of
20
ppb)
and
higher
than
MBA
concentrations
at
the
two
highest
influent
bromide
concentrations
(
ranging
from
50
to
>
100
ppb
(
Tables
IV­
5
and
IV­
7).

The
mean
concentrations
of
BCA
in
surface
water
were
statistically
significantly
higher
(
at
p
=
0.05)
than
the
mean
concentrations
of
BCA
in
groundwater
at
influent
bromide
concentrations
of
20
­
<
30
ppb
and
at
50
­
<
100
ppb.
There
were
no
statistically
significant
differences
(
at
p
=
0.05)
between
the
mean
concentrations
of
DBA
and
MBA
in
surface
water
and
the
mean
concentrations
of
these
chemicals
in
groundwater.

The
mean
concentrations
of
BCA
and
DBA
are
significantly
higher
in
surface
water
than
their
mean
concentrations
in
groundwater
for
a
given
influent
bromide
concentration
(
with
the
exception
of
DBA
at
influent
bromide
concentrations
at
<
20ppb).
However,
the
mean
MBA
concentrations
in
surface
water
were
significantly
higher
(
p
=
0.05)
than
the
mean
concentrations
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
23
Draft,
do
not
cite
or
quote
of
MBA
in
groundwater
only
for
the
influent
bromide
concentrations
ranging
from
20
ppb
to
<
50
ppb.

A.
2.3
Total
Organic
Carbon
(
TOC)
Concentration
in
ICR
Database
Many
researchers
have
documented
that
chlorine
reacts
with
natural
organic
matter
in
water
to
produce
a
variety
of
DBPs,
including
trihalomethanes
and
haloacetic
acids
(
Reckhow
and
Singer,
1990;
Reckhow
et
al.,
1990;
Marhaba
and
Van,
2000).
Natural
organic
matter
in
source
water
is
generally
monitored
as
total
organic
carbon
(
TOC).
Arora
et
al.
(
1997)
analyzed
results
of
a
DBP
survey
and
a
two­
year
DBP­
monitoring
study
of
more
than
100
treatment
plants
of
the
American
Water
System
from
1989
to
1991,
and
reported
no
correlation
between
rawwater
TOC
and
the
total
of
5
haloacetic
acid
(
HAA5)
concentrations
in
finished
and
distributedwater
samples.
A
significant
correlation
(
p
<
0.01)
was
found
between
TOC
and
HAA5
in
plant
effluent
and
distributed
water
samples.
However,
only
11
and
15
percent
of
the
variation
in
HAA5
was
explained
by
TOC
for
the
distributed­
water
samples
and
plant
effluent,
respectively.
No
published
studies
were
located
that
examined
the
effect
of
TOC
on
the
concentration
of
brominated
acids.

Tables
IV­
8,
IV­
9
and
IV­
10
present
data
from
the
ICR
database
for
the
concentrations
of
MBA,
BCA
and
DBA,
respectively,
as
a
function
of
influent
TOC
concentrations.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
24
Draft,
do
not
cite
or
quote
In
contrast
to
the
data
presented
by
Arora
(
1997),
a
regression
analysis
of
the
ICR
data
indicates
that
there
was
a
significant
correlation
(
 
=
0.05)
between
influent
TOC
concentration
and
the
mean
concentrations
of
MBA,
BCA,
and
DBA
in
surface
water
(
Tabes
IV­
5,
IV­
6,
and
IV­
7)/
This
correlation
with
TOC
levels
is
consistent
with
the
formation
of
brominated
acetic
acids
from
the
reaction
of
humic
acid
and
hypobromous
acid,
a
compound
formed
by
the
reaction
of
bromide
ion
with
ozone
and/
or
chlorine
in
the
disinfection
process
(
Chapter
II).
No
such
correlation
was
observed
in
groundwater,
which
had
lower
overall
TOC
levels.

Examination
of
the
data
using
the
Student's
t­
test
indicates
that,
with
a
few
exceptions,
at
a
given
TOC
concentration
in
surface
water
and
groundwater,
the
mean
concentrations
of
BCA
were
significantly
higher
(
p
=
0.05)
than
the
mean
concentrations
of
DBA,
which
were
significantly
higher
(
p
=
0.05)
than
the
mean
MBA
concentrations
(
Tables
IV­
8,
IV­
9,
and
IV­

10).
ion.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
25
Draft,
do
not
cite
or
quote
Table
IV­
8.
MBA
by
Influent
Total
Organic
Carbon
(
TOC)
Concentration
(
Quarterly
Distribution
System
Average)

Source
Influent
TOC
Conc.
(
ppb)
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
<
1
12
61
98.36
0.00
0.00
0.04
ND
0.28
0.00
0.00
1
­
<
2
58
271
95.57
0.04
0.00
0.26
ND
3.20
0.00
0.00
2
­
<
3
100
479
84.13
0.22
0.00
0.72
ND
6.63
0.00
0.83
3
­
<
4
60
306
76.14
0.28
0.00
0.61
ND
3.33
0.00
1.33

4
69
324
74.69
0.50
0.00
1.75
ND
26.00
0.00
1.55
GW
<
1
83
405
89.38
0.14
0.00
0.55
ND
6.68
0.00
0.28
1
­
<
2
13
52
65.38
0.38
0.00
0.65
ND
2.18
0.00
1.45
2
­
<
3
8
37
81.08
0.30
0.00
0.74
ND
3.08
0.00
1.23
3
­
<
4
3
7
85.71
0.10
0.00
0.26
ND
0.70
0.00
0.70

4
16
80
88.75
0.06
0.00
0.24
ND
1.95
0.00
0.28
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
ND:
Nondetect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
26
Draft,
do
not
cite
or
quote
Table
IV­
9.
BCA
by
Influent
Total
Organic
Carbon
(
TOC)
Concentration
(
Quarterly
Distribution
System
Average)

Source
Influent
TOC
Conc.
(
ppb)
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
<
1
12
62
53.23
0.74
0.00
1.00
ND
4.10
0.00
2.25
1
­
<
2
58
270
15.19
2.52
1.98
2.41
ND
15.75
0.00
5.25
2
­
<
3
100
481
8.11
3.59
2.80
3.07
ND
23.80
0.35
7.73
3
­
<
4
60
309
4.21
4.17
3.38
3.14
ND
22.50
1.10
8.18

4
69
323
3.10
4.67
3.98
3.19
ND
24.18
1.48
8.53
GW
<
1
83
405
63.46
0.65
0.00
1.40
ND
10.70
0.00
2.15
1
­
<
2
13
55
10.91
3.06
2.55
1.98
ND
7.35
0.00
5.93
2
­
<
3
8
36
38.89
2.20
1.39
2.43
ND
7.83
0.00
5.68
3
­
<
4
3
8
0.00
3.40
2.86
1.27
1.78
5.08
1.78
5.08

4
16
80
8.75
3.99
3.55
2.51
ND
11.28
0.43
7.64
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
ND:
Nondetect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
27
Draft,
do
not
cite
or
quote
Table
IV­
10.
DBA
by
Influent
Total
Organic
Carbon
(
TOC)
Concentration
(
Quarterly
Distribution
System
Average)

Source
Influent
TOC
Conc.
(
ppb)
Plants
N
PctND
%
Mean
µ
g/
L
Median
µ
g/
L
STD
µ
g/
L
Min
µ
g/
L
Max
µ
g/
L
p10
µ
g/
L
p90
µ
g/
L
SW
<
1
12
62
74.19
0.45
0.00
0.86
ND
2.68
0.00
2.08
1
­
<
2
58
272
72.79
0.46
0.00
1.10
ND
8.48
0.00
1.70
2
­
<
3
100
486
62.76
0.87
0.00
1.56
ND
7.65
0.00
3.25
3
­
<
4
60
309
55.02
1.69
0.00
2.81
ND
12.75
0.00
6.23

4
69
326
48.16
1.49
0.25
2.45
ND
14.25
0.00
4.33
GW
<
1
83
406
60.59
0.69
0.00
1.36
ND
13.00
0.00
2.28
1
­
<
2
13
54
31.48
1.54
1.11
1.71
ND
6.58
0.00
3.75
2
­
<
3
8
36
61.11
1.71
0.00
2.68
ND
7.25
0.00
6.73
3
­
<
4
3
8
100.00
0.00
0.00
0.00
ND
0.00
0.00
0.00

4
16
80
43.75
0.67
0.26
0.85
ND
3.48
0.00
2.14
1
Nondetects
are
treated
as
zero.
Source:
SW
­
Surface
Water,
GW
­
Groundwater
Plants:
Number
of
plants
sampled
N:
Number
of
samples
PctND:
Percent
samples
nondetect
Mean:
Arithmetic
mean
of
all
samples
Median:
Median
value
of
all
samples
STD:
Standard
deviation
Min:
Minimum
Value
Max:
Maximum
Value
p10:
10th
percentile
p90:
90th
percentile
ND:
Nondetect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
28
Draft,
do
not
cite
or
quote
A.
2.4
Seasonal
Shifts
Seasonal
shifts
in
brominated
acetic
acids
were
investigated
by
Krasner
et
al.
(
1989).
In
September
1987,
the
USEPA's
Office
of
Drinking
Water
entered
into
a
cooperative
agreement
with
the
Association
of
Metropolitan
Water
Agencies
(
AMWA)
to
perform
a
study
of
the
occurrence
and
control
of
DBPs.
The
AMWA
contracted
with
the
Metropolitan
Water
District
of
Southern
California
(
MWD)
to
provide
management
services
for
the
project
and
to
perform
the
DBP
analysis.
In
addition,
the
State
of
California
Department
of
Health
Services
(
CDHS),

through
the
California
Public
Health
Foundation
(
CPHF),
contracted
with
MWD
to
perform
a
similar
study
in
California.
Baseline
data
were
gathered
on
35
water­
treatment
facilities,
including
25
water
utilities
across
the
United
States
in
the
USEPA
study
and
10
California
water
utilities
in
the
CDHS
study.
Levels
of
MBA
and
DBA
were
measured,
but
BCA
was
not
evaluated.

During
the
first
quarter
(
spring
1988),
a
high
correlation
was
found
between
DBA
and
the
disinfectant
byproduct
dibromochloromethane.
In
addition,
Krasner
et
al.
(
1989)
reported
that
relatively
high
levels
of
the
measured
brominated
DBPs
were
detected
at
some
of
the
utilities.

These
findings
suggested
that
the
influence
of
bromide
in
the
raw
water
should
be
evaluated.

Therefore,
chloride
and
bromide
analyses
was
added
to
the
protocol,
beginning
with
the
second
quarter
(
summer
1988)
of
sampling.
Among
the
35
facilities,
bromide
levels
ranged
from
<
0.01
to
3.00
mg/
L.
At
the
utility
with
the
highest
bromide
levels
there
was
a
shift
in
the
distribution
of
DBPs
from
the
chlorinated
DBPs
to
the
brominated
DBPs,
resulting
in
DBA
as
the
major
haloacetic
acid
detected.
While
there
were
no
clear
trends
of
the
concentrations
of
ions
or
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
29
Draft,
do
not
cite
or
quote
brominated
acetic
acids
with
season
in
the
composite
analysis,
DBA
levels
increased
in
the
warmer
months
in
the
utility
with
the
highest
bromide
levels.
Some
observed
shifts
in
utilities
were
also
seen
as
the
result
of
drought
conditions
and
saltwater
intrusion.

A.
2.4.1
Seasonal
Shifts
in
ICR
Database
The
seasonal
mean
concentrations
of
MBA,
BCA,
and
DBA
are
presented
in
Tables
IV­

11,
IV­
12,
and
IV­
13,
respectively.
For
simplicity
of
presentation,
only
the
data
required
to
conduct
a
Student's
t­
test
has
been
presented
here.
Additional
data
can
be
located
in
the
ICR
database
(
U.
S.
EPA,
2000a).
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
30
Draft,
do
not
cite
or
quote
Table
IV­
11.
MBA
by
Sample
Quarter
(
Quarterly
Distribution
System
Average)

Sample
Quarter
MBA
Surface
Water
Ground
Water
N
Mean
(
µ
g/
L)
STD
(
µ
g/
L)
N
Mean
(
µ
g/
L)
STD
(
µ
g/
L)

Summer
`
97
239
0.29
0.73
92
0.21
0.79
Fall
`
97
250
0.23
0.55
86
0.21
0.60
Winter
`
98
240
0.25
1.74
101
0.18
0.53
Spring
`
98
262
0.17
0.53
105
0.14
0.46
Summer
`
98
250
0.28
0.83
104
0.09
0.30
Fall
`
98
229
0.32
1.03
93
0.13
0.51
N:
Number
of
samples
STD:
Standard
deviation
Sample
Quarter:

Summer
`
97:
July,
August,
and
September
Fall
`
97:
October,
November,
and
December
Winter
`
98:
January,
February,
and
March
Spring
`
98:
April,
May,
and
June
Summer
`
98:
July,
August,
and
September
Fall
`
98:
October,
November,
and
December
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
31
Draft,
do
not
cite
or
quote
Table
IV­
12.
BCA
by
Sample
Quarter
(
Quarterly
Distribution
System
Average)

Sample
Quarter
BCA
Surface
Water
Ground
Water
N
Mean
(
µ
g/
L)
STD
(
µ
g/
L)
N
Mean
(
µ
g/
L)
STD
(
µ
g/
L)

Summer
`
97
236
4.12
3.83
92
1.60
2.54
Fall
`
97
252
3.52
2.82
88
1.40
2.09
Winter
`
98
243
3.11
2.54
103
1.48
2.02
Spring
`
98
260
3.54
2.90
104
1.33
1.99
Summer
`
98
251
3.90
3.42
103
1.55
2.11
Fall
`
98
232
3.46
2.76
94
1.46
2.17
N:
Number
of
samples
STD:
Standard
deviation
Sample
Quarter:

Summer
`
97:
July,
August,
and
September
Fall
`
97:
October,
November,
and
December
Winter
`
98:
January,
February,
and
March
Spring
`
98:
April,
May,
and
June
Summer
`
98:
July,
August,
and
September
Fall
`
98:
October,
November,
and
December
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
32
Draft,
do
not
cite
or
quote
Table
IV­
13.
DBA
by
Sample
Quarter
(
Quarterly
Distribution
System
Average)

Sample
Quarter
DBA
Surface
Water
Ground
Water
N
Mean
(
µ
g/
L)
STD
(
µ
g/
L)
N
Mean
(
µ
g/
L)
STD
(
µ
g/
L)

Summer
`
97
241
1.30
2.19
92
0.89
1.92
Fall
`
97
252
1.12
2.08
88
0.88
1.39
Winter
`
98
244
0.94
2.04
103
0.81
1.33
Spring
`
98
262
0.88
1.86
104
0.67
1.30
Summer
`
98
253
1.17
2.16
104
0.80
1.53
Fall
`
98
232
1.14
2.16
93
0.88
1.41
N:
Number
of
samples
STD:
Standard
deviation
Sample
Quarter:

Summer
`
97:
July,
August,
and
September
Fall
`
97:
October,
November,
and
December
Winter
`
98:
January,
February,
and
March
Spring
`
98:
April,
May,
and
June
Summer
`
98:
July,
August,
and
September
Fall
`
98:
October,
November,
and
December
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
33
Draft,
do
not
cite
or
quote
An
examination
of
the
data
using
the
Student's
t­
test
showed
that,
based
on
only
two
seasons
of
analysis,
the
mean
concentrtions
of
MBA
in
surface
water
were
significantly
higher
(
p
=
0.05)
in
the
summer
than
in
the
spring
(
Table
IV­
11).
Also,
based
on
only
two
seasons
of
analysis,
the
mean
concentrations
of
BCA
in
surface
water
were
higher
in
summer
than
in
winter.

Aside
from
these
exceptions,
there
were
no
consistently
significant
differences
(
p
=
0.05)
in
the
mean
concentrations
of
BCA
or
DBA
between
one
season
and
other
in
surface
water
(
Table
IV­

13).
This
is
in
apparent
contrast
to
the
findings
of
Krasner
et
al.
(
1989),
who
found
that
DBA
levels
increased
in
the
warmer
months
in
the
utility
with
the
highest
bromide
levels.
Seasonal
variations
in
brominated
acetic
acids
may
be
dependent
on
seasonal
fluctuations
in
bromide­
ion
concentration,
which
were
not
evaluated
in
this
analysis.
No
seasonal
differences
in
mean
MBA,

BCA,
or
DBA
concentrations
in
groundwater
could
be
discerned.

B.
Exposure
to
Sources
Other
Than
Drinking
Water
MBA
has
been
used
in
industry
and
in
hospitals.
Between
1981
to
1983,
The
National
Institute
of
Occupational
Safety
(
NIOSH)
conducted
a
survey
of
a
sample
of
4490
businesses
employing
nearly
1,800,000
workers
(
NIOSH,
1990).
Potential
exposure
estimates
included
surveyor
observations
of
the
use
of
MBA
and
trade­
name
products
known
to
contain
MBA.

During
the
period
from
1981
to
1983,
4874
workers
were
potentially
exposed
to
MBA.
The
largest
numbers
of
exposures
(
1999)
occurred
in
the
commercial
printing
letterpress
business.

Used­
car
dealers
made
up
the
next
largest
number
of
potential
exposures
(
1723).
The
remainder
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
34
Draft,
do
not
cite
or
quote
of
potential
exposures,
in
decreasing
numbers,
included
workers
in
the
production
of
plastics
(
402),
in
hospitals
(
318),
and
in
individuals
working
with
electron
tubes
(
192).
Exposure
levels
were
not
reported
in
this
survey,
and
more­
recent
information
on
numbers
of
workers
exposed
was
not
available.

During
the
period
from
1981
to
1983,
there
were
no
reported
survey
observations
of
the
use
of
BCA,
DBA,
or
trade­
name
products
known
to
contain
BCA
or
DBA
in
the
workplace
(
NIOSH,
1990).

No
data
were
located
on
exposure
to
MBA,
BCA,
or
DBA
in
food,
air,
or
via
dermal
exposure
when
showering
or
swimming.

C.
Overall
Exposure
Only
limited
data
on
exposure
to
MBA,
BCA,
and
DBA
in
sources
other
than
drinking
water
exposure
was
located.
Exposure
to
drinking
water
is
discussed
in
Section
IV.
A.

D.
Body
Burden
No
data
could
be
located
on
body
burden.
However,
as
discussed
in
Chapter
3,
the
brominated
acetic
acids
are
metabolized
rapidly
and
are
not
lipophilic
at
physiological
pH,
and
so
would
not
be
expected
to
bioaccumulate.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
35
Draft,
do
not
cite
or
quote
E.
Summary
The
ICR
database
(
U.
S.
EPA,
2000a)
contains
extensive
information
on
concentrations
of
MBA,
BCA,
and
DBA
in
drinking­
water
systems,
and
on
how
those
concentrations
vary
with
input­
water
characteristics
and
treatment
methods.
The
database
contains
information
from
six
quarterly
samples
from
7/
97
to
12/
98,
from
approximately
300
large
systems
covering
approximately
500
plants.
The
mean
concentrations
of
BCA
were
1.47
and
3.61
µ
g/
L
from
groundwater
and
surface
water
respectively.
The
mean
concentrations
of
DBA
were
0.82
and
1.09
µ
g/
L
in
groundwater
and
surface
water,
respectively.
Examination
of
the
ICR
data
using
the
Student's
t­
test
indicates
that
the
mean
concentrations
of
MBA,
BCA,
and
DBA
in
surface
water
were
significantly
higher
(
p
=
0.05)
than
the
mean
concentrations
of
these
chemicals
in
groundwater.
In
addition,
the
mean
concentrations
of
BCA
were
significantly
higher
(
p
=
0.05)

than
the
mean
concentrations
of
DBA,
which
were
significantly
higher
(
p
=
0.05)
than
the
mean
MBA
concentrations
in
both
surface
water
and
groundwater.

Examination
of
the
ICR
data
using
the
Student's
t­
test
suggests
that,
although
the
concentrations
of
MBA
in
surface
water
treated
with
chlorine
are
similar
to
those
treated
with
chlorine
followed
by
chloramine.
BCA
and
DBA
concentrations
were
lower
when
free
chlorine
was
used
both
in
the
treatment
plant
and
the
distribution
system.
Although
ozonation
appeared
to
significantly
reduce
the
formation
of
BCA,
there
were
no
significant
differences
in
MBA
or
DBA
concentrations
(
p
=
0.05)
with
the
use
of
ozone
in
treating
surface
water
as
compared
to
the
common
(
non­
ozonation)
chemical­
disinfection
processes.
In
addition
there
were
no
significant
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
36
Draft,
do
not
cite
or
quote
differences
(
p
=
0.05)
between
the
two
treatments
using
ozonation
in
treating
surface
water
for
MBA,
BCA
and
DBA.

Consistent
with
the
findings
of
other
investigators,
and
the
chemistry
of
the
formation
of
bromoacetic
acids,
a
regression
analysis
of
the
ICR
data
indicates
that,
with
the
exception
of
MBA
in
surface
water,
there
was
a
significant
correlation
(
at
 
=
0.05)
between
influent
bromide
concentration
and
the
mean
concentrations
of
BCA
and
DBA
in
surface
water
and
groundwater.

In
addition,
for
a
given
influent
bromide
concentration
range,
the
mean
concentrations
of
BCA
were
generally
higher
(
p
=
0.05)
that
the
mean
concentrations
of
DBA
and
MBA
in
both
surface
water
and
groundwater.

A
regression
analysis
of
the
ICR
data
indicates
that
there
was
a
significant
correlation
(
 
=
0.05)
between
influent
TOC
concentration
and
the
mean
concentrations
of
MBA,
BCA,
and
DBA
in
surface
water.
This
is
consistent
with
the
formation
of
brominated
acetic
acids
from
the
reaction
of
humic
acid
and
hypobromous
acid,
a
compound
formed
by
the
reaction
of
bromide
ion
with
ozone
and/
or
chlorine
in
the
disinfection
process.
In
addition,
for
a
given
influent
TOC
concentration
range
in
surface
water,
the
mean
concentrations
of
BCA
were
significantly
higher
(
p
=
0.05)
than
the
mean
concentrations
of
DBA,
which
were
significantly
higher
(
p
=
0.05)
than
the
MBA
mean
concentrations.

Examination
of
the
data
using
the
Student's
t­
test
showed
that,
based
on
only
two
seasons
of
analysis,
the
mean
concentrations
of
MBA
in
surface
water
were
significantly
higher
(
p
=
0.05)

in
the
summer
than
in
the
spring
Also,
based
on
only
two
seasons
of
analysis,
mean
concentrations
of
BCA
in
surface
water
were
higher
in
summer
than
in
winter.
Aside
from
these
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IV­
37
Draft,
do
not
cite
or
quote
exceptions,
there
were
no
consistently
significant
differences
(
p
=
0.05)
in
the
mean
concentrations
of
MBA,
BCA
or
DBA
between
one
season
and
another
in
either
surface
water
or
groundwater.
Seasonal
variations
in
brominated
acetic
acids
may
be
dependent
on
seasonal
fluctuations
in
bromide­
ion
concentration,
which
were
not
evaluated
in
this
analysis.

The
data
on
exposure
to
sources
other
than
drinking
water
are
limited,
but
MBA
has
been
used
in
industry
and
in
hospitals.
Between
1981
to
1983,
4874
workers
were
potentially
exposed
to
MBA.
(
NIOSH,
1990).
No
data
were
located
on
exposure
to
MBA,
BCA,
or
DBA
in
food,
air,

or
via
dermal
exposure.

No
data
could
be
located
on
body
burden
levels
of
MBA,
BCA,
or
DBA..
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
1
Draft,
do
not
cite
or
quote
Chapter
V.
Health
Effects
in
Animals
The
available
database
for
the
brominated
acetic
acids
is
limited
and,
therefore,
many
toxicity
endpoints
have
not
been
fully
explored.
In
recognition
of
this
paucity
of
data,
there
is
a
large
body
of
ongoing
work,
particularly
for
BCA
and
DBA.
Preliminary
results
for
many
studies
have
been
reported
in
published
abstracts
and
are
included
here
to
provide
a
sense
of
the
spectrum
of
effects
induced
by
the
brominated
acetic
acids.
The
full
published
studies
would
need
to
be
evaluated
for
a
complete
understanding
of
the
chemicals'
effects
and
determination
of
the
relevance
of
the
data
for
quantitative
risk­
assessment
purposes.

A.
Short­
Term
Exposure
Monobromoacetic
acid
Linder
et
al.
(
1994a)
reported
on
the
acute
oral
toxicity
of
MBA
as
part
of
a
study
on
its
spermatogenic
effects.
Five
male
Sprague­
Dawley
rats
per
group
were
given
single
doses
of
100
to
200
mg/
kg
MBA
(
specific
dose
levels
not
reported)
by
gavage
in
water.
The
LD
50
was
reported
as
177
mg/
kg,
with
a
95%
(
fiducial)
confidence­
limit
range
of
156
to
226
mg/
kg.
Observed
clinical
symptoms
included
excess
drinking,
hypomobility,
labored
breathing,
and
diarrhea.
No
histopathologic
changes
were
observed
in
either
the
epididymal
sperm
or
testes
of
surviving
animals.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
2
Draft,
do
not
cite
or
quote
MBA
is
irritating
and
corrosive
to
the
human
skin
and
mucous
membranes
(
NIOSH,

2000).
The
ability
of
a
variety
of
carboxylic
acids
to
cause
skin
corrosion
was
investigated
in
support
of
the
development
of
a
multivariate
quantitative
structure­
activity
relationship
(
QSAR)

analysis
(
Eriksson
et
al.,
1994).
Forty­
five
aliphatic
carboxylic
acids
(
including
MBA,
DBA,
and
BCA)
were
evaluated
in
the
QSAR
analysis.
Fifteen
of
the
compounds,
including
MBA,

monochloroacetic
acid,
and
dichloroacetic
acid,
were
tested
for
cutaneous
corrosion
on
adult
rabbits.
In
these
studies,
the
test
chemical
was
applied
to
bare,
shaved
skin
under
an
occlusive
glass
filter
for
one
hour.
The
study
description
indicated
that
a
10
cm
by
10
cm
area
on
the
trunk
of
the
rabbit
was
shaved,
but
did
not
specify
that
this
entire
area
was
exposed.
The
observed
and
predicted
lowest­
observed­
effect
concentrations
(
LOECs)
for
inducing
corrosion
for
MBA
were
0.2
M
and
0.1
M,
respectively.

No
short­
term
toxicity
studies
for
exposure
by
the
inhalation
route
were
identified.

Bromochloroacetic
acid
Systemic
toxicity
was
evaluated
as
part
of
a
reproductive
and
developmental
toxicityscreening
assay
for
BCA
(
NTP,
1998).
As
part
of
a
range­
finding
study,
groups
of
male
and
female
Sprague­
Dawley
rats
(
6/
group)
were
exposed
for
14
days
to
drinking
water
containing
0,

30,
100,
300,
or
500
ppm
BCA.
The
study
authors
reported
that
the
estimated
average
doses
resulting
from
these
treatments
were
0,
3,
10,
28,
and
41
mg/
kg/
day.
No
mortality
was
observed
and
there
were
no
treatment­
related
differences
in
body
weight,
body­
weight
gain,
feed
and
water
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
3
Draft,
do
not
cite
or
quote
consumption,
or
clinical
observations
as
compared
with
controls.
The
NOAEL
for
general
toxicity
was
41
mg/
kg/
day.
A
LOAEL
could
not
be
determined.

The
results
of
the
14­
day
range­
finding
study
were
used
to
select
doses
for
a
reproductive
and
developmental
toxicity­
screening
assay
(
NTP,
1998).
Sprague­
Dawley
rats
were
administered
BCA
in
their
drinking
water
for
various
periods
during
a
35­
day
study
period.
Rats
were
divided
into
two
groups
of
males
and
three
groups
of
females.
Group
A
males
(
10/
group)
were
exposed
on
study
days
6
­
35
to
doses
of
0,
60,
200,
or
600
ppm.
Group
B
males
were
exposed
on
study
days
6
­
31
(
5/
group
at
0,
60,
or
200
ppm
and
8/
group
at
600
ppm
)
and
administered
bromodeoxyuridine
(
BrdU)
via
subcutaneously­
implanted
pumps
for
3
days
prior
to
necropsy
in
order
to
measure
cell
proliferation.
The
female
rats
were
grouped
as
follows:
Group
A
rats
(

periconception
exposure
group;
10/
dose
at
0,
60,
200,
or
600
ppm)
were
given
BCA
on
study
days
1
­
34
and
were
cohabitated
with
treated
males
on
study
days
13­
18.
Group
B
rats
(

gestationalexposure
group;
13/
dose)
were
cohabitated
with
treated
males
on
study
days
1
­
5
and
exposed
to
BCA
at
doses
of
0,
60,
200,
or
600
ppm
on
gestation
days
(
GD)
6
to
parturition.
Group
C
(

periconception
exposure
group;
5/
dose
group
at
0,
60,
200
ppm
and
8/
group
at
600
ppm)
were
exposed
to
BCA
on
study
days
1
­
30,
cohabitated
with
treated
males
on
study
days
13­
18,
and
administered
BrdU
via
subcutaneously
implanted
pumps
for
3
days
prior
to
necropsy
for
cellproliferation
assessment.

The
average
daily
doses
for
male
rats
in
both
Group
A
and
Group
B
were
estimated
by
the
study
authors
as
equivalent
to
0,
5,
15,
and
39
mg/
kg/
day.
No
mortality,
clinical
signs
of
toxicity,

or
body­
weight
changes
were
observed.
Water
consumption
was
decreased
by
21%
to
34%
at
the
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
4
Draft,
do
not
cite
or
quote
high
dose,
presumably
due
to
taste
aversion.
Clinical
chemistry
was
evaluated
in
Group
A
males,

and
the
following
statistically
significant
changes
were
observed:
a
16%
increase
in
albumin
to
globulin
ratio
in
high­
dose
animals;
decreased
alanine
aminotransferase
(
ALT)
activity
in
both
the
mid­
dose
(
15%
decrease)
and
high­
dose
(
20%
decrease)
groups;
and
increased
albumin
(
5%

increase
in
the
low­
dose
group,
no
significant
change
in
the
mid­
dose
group,
and
9%
increase
in
the
high­
dose
group).
Clinical­
chemistry
changes
of
these
magnitudes
are
not
generally
considered
to
be
toxicologically
significant
and
the
changes
in
the
high­
dose
group
may
have
been
secondary
to
dehydration.
Although
absolute
and
relative
liver
weights
increased
with
increasing
dose
in
both
Group
A
and
Group
B,
the
relative
liver
weight
was
statistically
different
from
controls
(
10%
increase)
only
in
the
high­
dose
group,
and
absolute
liver
weight
was
not
significantly
elevated.
Gross
necropsy
did
not
reveal
any
major
changes.
No
dose­
related
increases
in
individual
hepatocyte
necrosis
(
Group
A)
or
labeling
index
(
Group
B
only)
were
observed.

Histopathologic
examination
showed
an
increase
in
cytoplasmic
vacuolization
of
hepatocytes
of
treated
animals
in
Group
A.
This
effect
was
observed
in
all
dose
groups,
was
more
prominent
in
the
high­
dose
group,
and
was
absent
in
controls.
However,
cytoplasmic
vacuolization
was
observed
in
both
control
and
dosed
animals
of
Group
B,
and
was
not
increased
with
BCA
treatment.
Although
the
study
authors
suggested
that
the
biological
significance
of
these
changes
could
not
be
determined
without
evaluation
following
longer­
term
exposure,
the
highest
dose
was
considered
sufficient
to
induce
general
toxicity
under
the
conditions
of
this
study.
In
light
of
the
questions
concerning
the
biological
significance
of
the
clinical
chemistry,
liver
weight
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
5
Draft,
do
not
cite
or
quote
histopathology
changes,
39
mg/
kg/
day
(
the
highest
dose
tested)
was
considered
to
be
a
marginal
LOAEL
in
males,
and
the
corresponding
NOAEL
was
15
mg/
kg/
day.

The
estimated
average
daily
doses
for
Group
A
and
Group
C
females
(
peri­
conception
exposure
groups)
were
0,
6,
19,
and
50
mg/
kg/
day.
No
mortality,
clinical
signs
of
toxicity,
or
body­
weight
changes
were
observed
in
the
female
rats
of
these
groups,
but
water
consumption
was
decreased
by
24­
34%
at
the
high
dose;
these
findings
were
similar
to
those
observed
in
treated
male
rats.
Decreased
water
consumption
was
the
only
effect
observed
in
Group
A
females.

However,
Group
C
females
exhibited
a
dose­
related
increase
in
the
incidence
of
renal
tubular
dilatation/
degeneration
(
0/
5
in
controls,
2/
5
in
the
19
mg/
kg/
day
group,
and
3/
8
in
the
50
mg/
kg/
day
group).
A
statistical
analysis
of
these
data
indicated
that
the
incidences
in
treated
groups
were
not
statistically
different
from
those
in
the
controls
(
p>
0.05
using
the
Fisher
exact
test);
however,
sample
sizes
were
small
and,
thus,
the
power
of
the
analysis
was
limited.
No
detailed
pathology
report
was
available
to
evaluate
the
biological
significance
of
these
histopathological
changes
in
the
kidneys.
The
study
authors
concluded
that
only
the
high
dose
resulted
in
renal
toxicity.
Cell
proliferation
analysis
showed
small
differences
between
the
labeling
index
of
the
treated
and
control
groups
for
the
kidney
in
Group
B
males
and
for
the
liver
and
urinary
bladder
in
Group
C
females.
However,
these
changes
were
not
dose­
dependent
and
were
not
considered
to
be
biologically
significant.

Estimated
average
daily
doses
of
BCA
for
Group
B
females
(
gestational
exposure
group)

were
0,
10,
25,
and
61
mg/
kg/
day.
Similar
to
other
groups,
the
only
consistent
treatment­
related
effect
in
Group
B
females
was
decreased
water
consumption
at
the
high
dose.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
6
Draft,
do
not
cite
or
quote
Overall,
some
histopathology
findings
in
this
study
suggest
that
the
liver
(
as
evidenced
by
increased
relative
weight
and
increased
cytoplasmic
vacuolization
in
males)
and
kidney
(
as
indicated
by
increased
renal
tubular
dilatation/
degeneration
in
females)
may
be
target
organs
for
BCA
toxicity.
Although
liver
histopathology
was
only
observed
in
high­
dose
Group
A
males,

these
effects
were
considered
to
be
treatment­
related
by
the
study
authors,
yielding
a
marginal
LOAEL
of
39
mg/
kg/
day
for
equivocal
liver
effects.
The
study
authors
also
concluded
that
the
histopathological
changes
observed
in
the
kidneys
of
Group
C
females,
although
not
statistically
different
from
controls,
might
be
indicative
of
kidney
toxicity.
However,
kidney
histopathology
was
not
corroborated
by
the
kidney­
labeling
index,
suggesting
that
renal
toxicity
was
not
of
sufficient
severity
to
induce
cellular
proliferation
and
regeneration.
Based
on
the
study
authors'

interpretation
of
the
results,
39
mg/
kg/
day
was
selected
as
a
LOAEL,
and
the
corresponding
NOAEL
was
19
mg/
kg/
day.
Effects
on
reproductive
and
developmental
endpoints
and
determination
of
critical­
effect
levels
for
these
systems
are
described
in
Section
V.
C.

Parrish
et
al.
(
1996)
tested
whether
the
ability
of
brominated
acetic
acids
to
induce
oxidative
DNA
damage
was
due
to
peroxisome
proliferation.
(
This
study
is
described
more
fully
in
section
V.
D.)
Male
B6C3F1
mice
(
6/
dose
group)
were
administered
drinking
water
containing
0,
100,
500,
or
2000
mg/
L
BCA
for
3
weeks.
The
authors
did
not
provide
information
on
average
daily
doses.
Based
on
a
default
water­
intake
value
of
0.25
L/
kg/
day
for
male
B6C3F1
mice
(
U.
S.

EPA,
1988),
the
corresponding
doses
were
estimated
to
be
0,
25,
125,
and
500
mg/
kg/
day.

General
toxicity
was
assessed
by
measuring
body
weight
and
liver
weight;
these
data
for
BCA
are
summarized
in
Table
V­
1.
The
effects
of
BCA
on
oxidative
DNA
damage
and
peroxisome
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
7
Draft,
do
not
cite
or
quote
proliferation
were
measured
in
the
livers
of
male
B6C3F1
mice,
using
the
following
measures:
(
1)

changes
in
the
DNA
adduct
8­
hydroxy­
2­
deoxyguanosine
(
8­
OhdG)
as
an
indicator
of
oxidative
stress,
and(
2)
changes
in
levels
of
cyanide
insensitive
Acyl­
CoA
oxidase
and
12­
hydroxylation
of
lauric
acid
as
indicators
of
peroxisome
proliferation.
An
additional
dose
group
exposed
to
3000
mg/
L
BCA
(
750
mg/
kg/
day)
was
evaluated
for
the
Acyl­
CoA
activity.
Body
weight
was
decreased
by
8.5%
at
the
highest
dose
tested.
Absolute
and
relative
liver
weights
were
increased
at
the
high
dose
by
20%
and
33%,
respectively.
BCA
had
no
effect
on
either
measure
of
peroxisome
proliferation
after
exposures
up
to
3000
mg/
L.
BCA
did
induce
oxidative
DNA
damage,
with
8­
OHdG
levels
in
nuclear
DNA
of
the
liver
significantly
increased
(
p<
0.05)

beginning
at
the
lowest
dose,
25
mg/
kg/
day.
The
level
of
8­
OHdG
increased
to
a
maximum
of
approximately
2­
fold
at
the
highest
dose
(
500
mg/
kg/
day).

The
absence
of
other
measures
of
liver
toxicity,
such
as
histopathology
or
clinical
chemistry
results,
clouds
the
classification
of
these
liver­
weight
changes
as
adverse.
However,
the
accompanying
increase
in
oxidative
DNA
damage
suggests
potentially
adverse
liver
effects.
Based
on
the
data
presented,
the
LOAEL
for
minimal
liver
effects
was
500
mg/
kg/
day
and
the
NOAEL
was
125
mg/
kg/
day.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
8
Draft,
do
not
cite
or
quote
Table
V­
1.
Body
and
Liver
Weight
Changes
Induced
by
BCA
and
DBAa
BCA
Drinking
water
(
g/
L)
Body
weight
(
g)
Liver
weight
(
g)
Relative
liver
weight
(%
body
weight)

Control
27.2
±
0.5c
1.5
±
0.01
5.4
%
±
0.1
0.1
(
25
mg/
kg/
day)
b
26.1
±
0.3
1.5
±
0.1
5.8
%
±
0.4
0.5
(
125
mg/
kg/
day)
28.3
±
0.7
1.8
±
0.1
6.2
%
±
0.6
2.0
(
500
mg/
kg/
day)
24.9
±
0.7**
d
1.8
±
0.1*
7.2%
±
0.4**

DBA
Control
27.1
±
0.5
1.5
±
0.01
5.4
%
±
0.1
0.1
(
25
mg/
kg/
day)
24.0
±
0.7**
1.4
±
0.1
5.8%
±
0.4
0.5
(
125
mg/
kg/
day)
25.6
±
0.8
2.1
±
0.2**
8.0%
±
0.5**

2.0
(
500
mg/
kg/
day)
26.1
±
0.4
2.0
±
0.1**
7.8%
±
0.6
Notes:

a.
Adapted
from
Parrish
et
al.,
1996
b.
Estimated
daily
doses
were
calculated
based
on
default
drinking
water
values
of
0.25
c.
Mean
±
standard
error
*
Statistical
significance:
p<
0.05
**
Statistical
significance:
p<
0.01
The
ability
of
a
variety
of
carboxylic
acids
to
cause
skin
corrosion
was
investigated
using
multivariate
quantitative
structure­
activity
relationship
QSAR
(
Eriksson
et
al.,
1994).
The
predicted
lowest­
observed­
effect
concentration
(
LOEC)
for
corrosion
for
BCA
was
0.7
M.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
9
Draft,
do
not
cite
or
quote
No
short­
term
toxicity
studies
for
exposure
to
BCA
by
the
inhalation
route
were
identified.

Dibromoacetic
acid
The
acute
toxicity
of
DBA
was
examined
as
part
of
a
study
on
the
spermatogenic
effects
of
this
compound
(
Linder
et
al.,
1994a).
Male
Sprague­
Dawley
rats
(
5/
group)
were
given
single
doses
of
1000
to
2000
mg/
kg
DBA
(
specific
dose
levels
not
reported)
by
oral
gavage
in
water.

Custom­
synthesized
DBA
of
>
99%
purity
was
used
because
the
commercial
chemical
is
generally
contaminated
with
approximately
10%
MBA.
Surviving
animals
were
killed
14
­
21
days
after
dosing.
The
oral
LD
50
was
1737
mg/
kg,
with
a
95%
(
fiducial)
confidence­
limit
range
of
1411
­

1952
mg/
kg.
Most
of
the
animal
deaths
occurred
within
48
hours
of
dosing.
Observed
symptoms
included
excess
drinking,
hypomobility,
labored
breathing,
diarrhea,
and
ataxia.
Histopathologic
examination
of
the
epididymal
sperm
in
surviving
animals
showed
the
presence
of
mis­
shapen
and
degenerating
sperm,
as
well
as
abnormal
retention
of
Step
19
spermatids.
Effects
other
than
spermatotoxicity
were
not
examined.

In
another
single­
dose
spermatotoxicity
study,
Vetter
et
al.
(
1998)
treated
sexually­
mature
male
Crl:
CD(
SD)
BR
rats
(
4­
5/
dose
group)
with
0,
600,
or
1200
mg/
kg
DBA
in
10
mL/
kg
deionized
water.
In
the
high­
dose
group,
signs
of
overt
toxicity
included
lethargy,
irregular
gait,

decreased
feces,
ocular
discharge,
and
dyspnea.
Abnormal
respiratory
sounds
were
observed
in
some
animals
(
number
of
animals
affected
was
not
specified)
and
one
animal
in
the
high­
dose
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
10
Draft,
do
not
cite
or
quote
group
died
on
study
day
3.
No
overt
toxicity
was
observed
in
the
low­
dose
group.
No
changes
in
measured
sperm
parameters
(
motility,
morphology,
and
cell­
membrane
permeability)
were
reported
at
either
dose;
however,
mi
testes
histopathology
(
the
presence
of
basophilic
bodies)
was
observed
in
both
dose
groups.
Based
on
the
clinical
findings,
1200
mg/
kg
was
considered
to
be
an
acute
frank
effects
level
(
FEL).
The
LOAEL
was
600
mg/
kg
for
testes
histopathology.
and
a
A
NOAEL
could
not
be
determined.

Parrish
et
al.
(
1996)
tested
whether
the
ability
of
brominated
acetic
acids
to
induce
oxidative
DNA
damage
was
due
to
peroxisome
proliferation.
(
This
study
is
described
more
fully
in
section
V.
D).
Male
B6C3F1
mice
(
6/
dose
group)
were
given
drinking
water
containing
0,
100,

500,
or
2000
mg/
L
DBA
for
3
weeks.
The
authors
did
not
provide
information
on
average
daily
doses.
However,
based
on
a
default
water­
intake
value
of
0.25
L/
kg/
day
for
male
B6C3F1
mice
(
U.
S.
EPA,
1988),
the
corresponding
daily
doses
were
estimated
to
be
0,
25,
125,
and
500
mg/
kg/
day,
respectively.
The
effects
of
DBA
on
oxidative
DNA
damage
and
peroxisome
proliferation
were
measured
in
the
livers
of
male
B6C3F1
mice,
using
the
following
measures:
(
1)

changes
in
the
DNA
adduct
8­
hydroxy­
2­
deoxyguanosine
(
8­
OhdG)
as
an
indicator
of
oxidative
stress,
and
(
2)
changes
in
levels
of
cyanide
insensitive
Acyl­
CoA
oxidase
and
12­
hydroxylation
of
lauric
acid
as
indicators
of
peroxisome
proliferation.
An
additional
dose
group
exposed
to
3000
mg/
L
DBA
(
750
mg/
kg/
day)
was
evaluated
for
the
Acyl­
CoA
activity
.
As
part
of
this
study,

general
toxicity
was
assessed
by
measuring
body
weight
and
liver
weight,
as
summarized
in
Table
V­
1.
No
dose­
related
decrease
in
body
weight
was
observed,
but
absolute
and
relative
liver
weights
were
increased
at
the
mid­
and
high­
doses.
At
the
mid­
dose
(
125
mg/
kg/
day),
absolute
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
11
Draft,
do
not
cite
or
quote
and
relative
liver
weights
were
increased
by
40%
and
48%,
respectively.
At
the
high
dose
(
500
mg/
kg/
day),
absolute
and
relative
liver
weights
were
increased
by
33%
and
44%,
respectively.

DBA
induced
Acyl­
CoA
activity
to
a
maximum
of
3­
fold
following
exposures
up
to
3000
mg/
L,

but
did
not
induce
the
12­
hydroxylation
of
lauric
acid.
DBA
induced
oxidative
DNA
damage,

with
8­
OHdG
levels
in
hepatic
nuclear
DNA
significantly
increased
(
p<
0.05)
at
the
highest
dose
(
500
mg/
kg/
day)
to
a
maximum
of
approximately
twice
the
control
response.
The
absence
of
a
clear
dose
response,
and
the
lack
of
other
measures
of
liver
toxicity,
such
as
histopathology
or
clinical
chemistry,
clouds
the
classification
of
the
liver­
weight
changes
as
adverse.
However,
the
magnitude
of
the
change
and
the
accompanying
increase
in
oxidative
DNA
damage
suggests
potentially­
adverse
liver
effects.
Based
on
the
data
presented,
the
LOAEL
for
minimal
liver
effects
was
125
mg/
kg/
day.
The
NOAEL
was
25
mg/
kg/
day.

As
part
of
a
male
reproductive
study,
Linder
et
al.
(
1995)
administered
daily
gavage
doses
of
0
or
250
mg/
kg
DBA
to
male
Sprague­
Dawley
rats
(
10/
dose
group).
Dosing
was
terminated
after
42
days
because
severe
toxic
effects,
including
labored
breathing,
light
tremor,
difficulty
moving
the
hind
limbs,
and
significant
weight
loss,
developed.
In
a
subsequent
study,
male
rats
(
10/
dose
group)
were
given
0,
2,
10,
or
50
mg/
kg/
day
DBA
by
oral
gavage
for
31
or
79
days.
The
only
observed
effect
was
a
slight
decrease
in
body
weight
in
the
50
mg/
kg/
day
group
on
day
79
to
approximately
95%
of
controls.
Based
on
the
findings
of
both
studies,
the
FEL
for
general
toxicity
was
250
mg/
kg/
day
and
the
NOAEL
was
50
mg/
kg/
day.
Adverse­
effect
levels
for
reproductive
endpoints
are
described
in
Section
V.
C.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
12
Draft,
do
not
cite
or
quote
NTP
(
1999)
evaluated
the
immunotoxicity
of
DBA
in
female
B6C3F1
mice
(
8/
dose
group)

exposed
to
drinking
water
containing
0,
125,
250,
500,
1000,
or
2000
mg/
L
DBA
for
28
days.

Four
separate
studies
were
conducted
and
different
general
and
immunologic
endpoints
were
examined
in
each
study.
Studies
1­
3
investigated
selected
immunologic
endpoints
and
bodyweight
changes;
body
and
organ
weights,
hematology,
and
gross
pathology
were
examined
in
Study
4.
Key
immunotoxicity
responses
are
presented
in
Table
V­
2,
and
body
and
organ
weights
are
summarized
in
Table
V­
3.
The
study
authors
did
not
estimate
DBA
daily
doses
resulting
from
drinking­
water
exposures;
however,
the
average
daily
doses
could
be
calculated
based
on
waterconsumption
and
body­
weight
data
provided
in
the
study
report.
DBA
dose
ranges
were
similar
across
the
four
studies
and
are
presented
below
in
conjunction
with
the
experimental
findings
for
each
study.
No
significant
differences
(
p<
0.05)
in
drinking­
water
consumption
among
dosed
groups
and
no
clinical
signs
of
overt
toxicity
were
observed
in
any
of
the
studies.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
13
Draft,
do
not
cite
or
quote
Table
V­
2.
Immunotoxicity
of
DBA
in
Female
B6C3F1
mice
a
Study
Number/
Endpoint
Dose
(
mg/
kg/
day)
b
Study
#
1
0
19
39
73
150
285
Spleen
cell
number
x
107
15.97
±
0.50c
16.69
±
0.51
17.55
±
1.02
19.13
±
0.58d**
18.49
±
0.46*
18.45
±
0.39*

Spleen
Macrophages
(%
of
cells)
2.8
±
0.2
3.0
±
0.2
2.9
±
0.2
3.5
±
0.3
4.2
±
0.1**
4.5
±
0.2**

Natural
Killer
cell
lytic
activity
(
LU/
10
7
cells)
e
12
±
1
14
±
2
15
±
1
16
±
1*
20
±
1**
24
±
2**

Mixed
leukocyte
response
(
CPM/
105
spleen
cells
­

responders)
633
±
69
834
±
160
863
±
126
781
±
52
693
±
50
851
±
69
Study
#
2
0
20
38
70
143
280
IgM
antibody­
forming
colonies
(
AFC/
106
spleen
cells)
1958
±
145
1767
±
62
1590
±
149
1333
±
89**
1251
±
120**
985
±
69**

Serum
IgM
titer
(
OD)
g
103
±
7
100
±
9
131
±
19
102
±
12
97
±
5
85
±
5
Study
#
4
0
14
33
68
132
236
Macrophage
Activating
Factor
(%
suppression
without
stimulation)

(%
suppression
with
stimulation)
29.08
±
2.77
86.80
±
4.26
19.61
±
5.80
92.02
±
2.04
16.45
±
6.39
94.17
±
2.16
10.22
±
5.86*

92.01
±
4.01
15.39
±
2.50
93.62
±
2.84
28.58
±
3.62
89.05
±
2.73
a.
Adapted
from
NTP,
1999
b.
Doses
were
estimated
based
on
drinking­
water
concentrations
and
water­
intake
and
body­
weight
data
provided
c.
Lytic
unit
(
LU)
=
the
number
of
splenocytes
required
to
kill
10%
of
the
target
cells.

d.
CPM
=
counts
per
minute
based
on
3H­
thymidine
incorporation
in
responder
cells.

e.
IgM
titer
based
on
enzyme­
linked
immunosorbant
assay.

*
Statistical
significance:
p<
0.05
**
Statistical
significance:
p<
0.01
Drinking
Water
Criteria
Document
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Acids
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HECD
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14
Draft,
do
not
cite
or
quote
Table
V­
3.
General
toxicity
of
DBA
in
Female
B6C3F1
micea
Estimated
Dose
(
mg/
kg/
day)
b
0
14
33
68
132
236
Body
weight
(
g)
23.7
±
0.4
24.0
±
0.4
24.7
±
0.6
24.8
±
0.7
24.6
±
0.4
23.0
±
0.5
Liver
weight
(
mg)
(%
of
body
weight)
1071
±
25
4.5
±
0.1
1183
±
32*
4.9
±
0.1*
1293
±
41**
5.2
±
0.1**
1386
±
62**
5.6
±
0.1
1479
±
14**
6.0
±
0.1**
1567
±
41**
6.8
±
0.1**

Kidney
weight
(
mg)
(%
of
body
weight)
287
±
15
1.21
±
0.07
298
±
6
1.24
±
0.01
305
±
5*
1.23
±
0.02
312
±
9*
1.26
±
0.02
317
±
6
*
1.29
±
0.01*
335
±
6*
1.46
±
0.01a*

Spleen
weight
(
mg)
(%
body
weight)
77
±
3
0.328
±
0.015
84
±
3
0.349
±
0.013
86
±
4
0.350
±
0.018
92
±
4*
0.370
±
0.015
88
±
3
0.359
±
0.011
96
±
4**
0.418
±
0.014**

Thymus
weight
(
mg)
(%
body
weight)
67
±
4
0.280
±
0.014
60
±
4
0.251
±
0.017
65
±
3
0.265
±
0.010
63
±
4
0.252
±
0.011
58
±
2
0.236
±
0.007
43
±
3**
0.189
±
0.013**

Reticulocyte
count
(%)
4.01
±
0.11
4.24
±
0.27
4.70
±
0.49
4.61
±
0.16*
4.50
±
0.16*
5.21
±
0.25**

Notes:

a.
Adapted
from
NTP,
1999
b.
Doses
were
estimated
based
on
drinking­
water
concentrations
and
water­
intake
and
body­
weight
data
provided
in
the
report.

c.
Mean
±
Standard
Error.

*
Statistical
significance:
p<
0.05
**
Statistical
significance:
p<
0.01
Drinking
Water
Criteria
Document
for
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Acetic
Acids
EPA/
OW/
OST/
HECD
V­
15
Draft,
do
not
cite
or
quote
In
Study
1,
doses
estimated
using
group­
specific
body
weights
and
water­
consumption
rates
were
0,
19,
39,
73,
150,
and
285
mg/
kg/
day.
A
statistically
significant
8%
decrease
in
terminal
body
weight
was
observed
in
the
high­
dose
group.
Spleen­
cell
number
was
significantly
increased
above
controls
(
p<
0.05)
at
73
mg/
kg/
day
and
higher,
but
there
was
no
dose­
response.

At
73
mg/
kg/
day,
spleen­
cell
number
was
elevated
approximately
20%
above
controls,
whereas
at
both
150
and
285
mg/
kg/
day,
the
increase
was
approximately
12%.
With
the
notable
exception
of
macrophages,
the
increase
in
the
absolute
number
of
most
spleen­
cell
types
paralleled
the
increase
in
total
number
of
spleen
cells;
thus
the
percentage
of
each
cell
type
was
generally
unaltered.
In
contrast,
spleen
macrophages
increased
in
a
dose­
dependent
manner
to
50%,
77%,
and
91%

above
controls
at
73,
150,
and
285
mg/
kg/
day,
respectively,
(
500,
1000,
and
2000
mg/
L,

respectively)
An
increase
in
natural
killer
(
NK)
cell
lytic
activity
was
also
observed
at
the
three
highest
DBA
dose
groups
(
p<
0.05)
when
expressed
as
specific
activity;
a
significant
increase
was
observed
in
the
four
highest
dose
groups
(
i.
e.,
39
mg/
kg/
day
and
above)
when
expressed
as
totalspleen
activity.
NK­
cell
activity
was
maximal
at
the
highest
dose,
showing
an
increase
of
100%

when
measured
as
specific
activity
and
of
143%
when
measured
as
total­
spleen
activity.
However,

the
significance
of
the
treatment­
related
changes
in
NK­
cell
activity
is
unclear,
because
the
positive
control
used
in
the
experiment
produced
a
significant
decrease
in
NK­
cell
activity.
DBA
treatment
had
no
effect
on
mixed­
leukocyte
response,
which
measures
proliferative
response
of
splenic
leukocytes
from
treated
animals
to
allogenic
lymphocytes
(
i.
e.,
lymphocytic
cells
from
a
genetically
distinct
strain
of
the
same
species)
of
DBA/
2
mice.
Overall,
the
results
of
Study
1
Drinking
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Acids
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OW/
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HECD
V­
16
Draft,
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or
quote
demonstrated
an
increase
in
several
measures
of
cellular
immunity,
with
statistically
significant
effects
generally
occurring
at
doses
of
73
mg/
kg/
day
and
higher.
The
toxicological
significance
of
the
findings
of
Study
1
are
unclear
and
are
discussed
in
more
detail
later
in
this
section,
using
a
weight­
of­
evidence
based
on
the
results
of
all
four
studies.

In
Study
2,
estimated
average
daily
doses
were
0,
20,
38,
70,
143,
and
280
mg/
kg/
day.
No
significant
effects
of
DBA
treatment
on
body
weight
were
observed.
In
contrast
to
Study
1,
no
increase
in
spleen­
cell
number
was
observed.
A
statistically
significant
(
p<
0.05)
dose­
dependent
decrease
in
spleen
IgM
antibody­
forming
cell
response
to
sheep
erythrocytes
was
observed
beginning

70
mg/
kg/
day.
There
was
no
change,
however,
in
serum­
IgM
titer
to
sheep
erythrocytes.
The
study
authors
noted
that
the
lack
of
concordance
between
these
two
measures
of
humoral
response
was
not
uncommon.
They
suggested
that
discordance
might
arise
from
the
fact
that
the
IgM
antibody­
forming
cell
response
is
a
specific
measure
of
immune
response
in
the
spleen,
whereas
the
IgM
titer
measures
systemic
humoral
immunity
and,
thus,
would
reflect
changes
in
both
bone­
marrow
and
lymph­
node
antibody
production,
in
addition
to
antibody
production
in
the
spleen.
Therefore,
the
IgM
assay
might
not
be
a
sensitive
measure
of
toxicity
for
substances
whose
target
organ
of
toxicity
is
only
the
spleen.

Study
3
evaluated
macrophage
activation.
The
estimated
average
daily
doses
were
0,
16,

35,
69,
134,
and
229
mg/
kg/
day.
No
significant
effects
on
body
weight
were
observed.
To
assess
macrophage
activation,
peritoneal
macrophages
were
stimulated
by
treatment
with
a
combination
of
gamma
interferon
and
lipopolysaccharide,
and
their
ability
to
kill
or
inhibit
the
growth
of
Drinking
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Document
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HECD
V­
17
Draft,
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not
cite
or
quote
B16F10
tumor
cells
was
measured.
No
clear
dose­
dependent
effects
of
DBA
treatment
on
B16F10
tumor­
cell
growth
were
observed
with
or
without
macrophage
activation.

Body
weight,
organ
weight,
gross
pathology,
and
hematology
were
evaluated
in
Study
4
(
Table
V­
3).
The
estimated
average
daily
doses
were
0,
14,
33,
68,
132,
and
236
mg/
kg/
day.

Based
on
pooled
body­
weight
data
for
all
studies,
no
significant
change
in
terminal
body
weight
was
observed.
However,
body­
weight
gain
was
decreased
by
40%
at
the
high
dose.
Statistically
significant
changes
in
organ
weight
were
also
reported.
Thymus
weight
was
significantly
decreased
only
at
the
high
dose.
Absolute
spleen
weight
was
elevated
at
all
the
doses,
but
was
statistically
different
from
controls
(
p<
0.05)
only
at
68
mg/
kg/
day
and
236
mg/
kg/
day
(
19%
and
24%
increase,
respectively,
as
compared
with
controls).
Although
absolute
spleen
weight
was
also
elevated
at
132
mg/
kg/
day
(
14%
increase
relative
to
controls),
this
difference
was
not
statistically
significant,
indicating
an
ambiguous
dose­
response.
Relative
spleen
weight
was
statistically
significantly
increased
only
at
the
high
dose,
although
there
was
an
increasing
trend
with
increasing
dose.
The
absolute
and
relative
liver­
weight
increases
were
dose­
dependent
and
were
significantly
elevated
at
all
doses
tested
(

14/
mg/
kg/
day).
Relative
liver
weight
increased
with
increasing
dose
to
9%,
16%,
24%,
33%,
and
51%
above
control
values
for
each
of
the
dose
groups.
Kidney
weight
was
also
statistically
increased
in
a
dose­
dependent
manner,
beginning
at
33
mg/
kg/
day
for
absolute
weight
(
6%
increase)
and
132
mg/
kg/
day
for
relative
kidney
weight
(
7%
increase).
With
the
exception
of
a
dose­
related
increase
in
reticulocytes
that
achieved
statistical
significance
at

68
mg/
kg/
day,
no
biologically
significant
hematology
parameters
were
Drinking
Water
Criteria
Document
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Acetic
Acids
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HECD
V­
18
Draft,
do
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or
quote
affected
by
treatment.
It
should
be
noted,
however,
that
increases
in
reticulocytes
are
generally
not
considered
adverse.
No
gross
pathological
lesions
were
identified.
Although
a
histopathologic
examination
was
conducted
for
Study
4,
no
results
were
provided.
Therefore,
treatment
histopathology
could
not
be
assessed.

Overall,
exposure
to
DBA
in
drinking
water
for
28
days
resulted
in
body­
and
organweight
changes
and
alterations
in
several
indicators
of
immunologic
response.
General
toxicity
indicators
included
decreased
body­
weight
gain
in
the
highest­
dose
group
tested
in
all
four
studies,
increased
liver
weights
(
both
absolute
and
relative)
at
all
doses
tested

14
mg/
kg/
day),

and
increased
absolute
(
33
mg/
kg/
day
and
above)
and
relative
(
132
mg/
kg/
day
and
above)
kidney
weights.
However,
the
absence
of
supporting
clinical
chemistry
and/
or
histopathologic
data
precludes
determination
of
liver
and
kidney
effects
as
adverse.
Therefore,
the
observed
increase
in
liver
weight
was
not
selected
as
the
critical
effect.
In
Study
4,
thymus
weights
were
decreased
and
spleen
weights
(
absolute
and
relative)
were
increased
at
the
highest
dose
tested
(
236
mg/
kg/
day).

Spleen
weight
was
also
increased
in
a
dose­
dependent
manner
at
lower
doses;
however,
the
doseresponse
was
not
clear
cut.
Further,
spleen
weights
were
not
increased
in
Study
2
at
similar
doses.

Therefore,
the
toxicologic
significance
of
this
finding
remains
unclear.
A
number
of
measures
of
cellular
and
humoral
immunity
were
altered
in
a
dose­
dependent
manner
by
DBA
treatment,

beginning
in
the
animals
treated
with
500
mg/
L
DBA
in
drinking
water
(
equivalent
to
68­
73
mg/
kg/
day,
depending
on
the
study).
Spleen­
cell
number
was
increased
in
Study
1
but
was
not
elevated
in
Study
2,
limiting
interpretation
of
these
findings.
As
previously
discussed,
NK­
cell
Drinking
Water
Criteria
Document
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Brominated
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Acids
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OW/
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HECD
V­
19
Draft,
do
not
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or
quote
lytic
activity
(
expressed
as
either
specific
activity
or
total
spleen
activity)
was
increased
in
DBAtreated
animals
in
Study
2,
but
was
decreased
by
exposure
to
a
positive
control
in
the
same
experiment.
Therefore,
these
findings
are
inconclusive.
The
number
of
spleen
macrophages,

however,
increased
statistically
in
a
dose­
dependent
manner
a

500
mg/
L
(
73
mg/
kg/
day)
DBA,

indicating
an
immunotoxic
response
in
this
target
organ.
In
Study
3,
exposure
to
500
mg/
L
(
70
mg/
kg/
day)
DBA
and
above
also
decreased
spleen
IgM
antibody­
forming
cell
response,
which
represents
a
depression
in
humoral
immunity.
The
decrease
in
spleen
IgM
antibody­
forming
cell
response
and
the
increase
in
spleen
macrophages
are
the
clearest
indicators
of
an
immunotoxic
effect.
Based
on
these
findings,
the
LOAEL
for
immunotoxicity
under
the
conditions
of
this
study
was
70
mg/
kg/
day,
and
the
NOAEL
was
38
mg/
kg/
day.

No
short­
term
toxicity
studies
for
exposure
to
DBA
by
the
inhalation
or
dermal
route
were
identified.

B.
Long­
Term
Exposure
Monobromoacetic
acid
No
long­
term
toxicity
studies
for
any
exposure
route
were
identified.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
20
Draft,
do
not
cite
or
quote
Bromochloroacetic
acid
In
a
published
abstract,
Stauber
et
al.
(
1995)
reported
on
preliminary
data
suggesting
that
BCA
induces
liver
tumors
in
B6C3F1
mice.
Noncancer
liver
effects
included
glycogen
accumulation
and
hepatocyte
vacuolization.
No
long­
term
toxicity
studies
for
any
exposure
route
were
identified
in
the
peer­
reviewed
literature.
However,
BCA
is
currently
undergoing
90­
day
subchronic
and
2­
year
chronic
bioassays
(
NTP,
2000a).

Dibromoacetic
acid
In
published
abstracts,
So
and
Bull
(
1995)
reported
that
DBA
increased
the
formation
of
aberrant
crypt
foci
in
the
colon
of
treated
rats,
and
Stauber
et
al.
(
1995)
reported
on
preliminary
data
suggesting
that
DBA
induces
liver
tumors
in
B6C3F1
mice.
The
effects
of
longer­
term
DBA
exposure
on
noncancer
endpoints
were
not
described.

Phillips
et
al.
(
2002,
published
abstract)
examined
the
neurotoxicity
of
DBA
in
adolescent
(
28­
day­
old)
male
and
female
F344
rats
(
12/
sex/
dose)
given
DBA
in
drinking
water
at
concentrations
of
0,
200,
600,
or
1500
mg/
L
(
mean
doses
calculated
by
the
authors
as
0,
20,
72,

and
161
mg/
kg/
day)
for
6
months.
In
both
sexes,
body
weight
was
significantly
depressed
in
the
high­
dose
group
but
overall
health
status
was
unaltered.
A
neurobehavioral
test
battery
was
administered
to
all
animals
at
1,
2,
4,
and
6
months.
Dose­
dependent
neuromuscular
toxicity,
Drinking
Water
Criteria
Document
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HECD
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Draft,
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or
quote
characterized
by
mild
gait
abnormalities,
hypotonia,
and
decreased
forelimb
and
hindlimb
grip
strength,
was
observed
in
both
sexes.
Sensorimotor
responsiveness,
as
measured
by
responses
to
a
tail
pinch
and
click,
was
reduced
at
all
doses,
but
did
not
progress
with
continued
exposure
to
DBA.
Decreased
motor
activity
was
noted
in
both
sexes
in
the
high­
dose
group,
whereas
a
chest
clasping
response
was
only
observed
in
high­
dose
females.
Neuropathologic
examination
revealed
significant
myelin
sheath
degeneration,
axonal
swelling,
and
axonal
degeneration
in
the
lateral
and
ventral
areas
of
the
spinal
cord
white
matter
in
the
high­
dose
group.
In
the
mid­
and
high­
dose
groups,
small
numbers
of
swollen,
eosinophilic
or
faintly
basophilic,
and
occasionally
vacuolated
neurons
were
observed
in
the
spinal
cord
gray
matter,
and
appeared
to
represent
axonal
degeneration.
Neuropathologic
examination
has
not
yet
been
conducted
in
the
low­
dose
group.

No
treatment­
related
neuropathology
was
noted
in
the
eyes,
peripheral
nerves,
peripheral
ganglia,

or
brain.
Based
on
neurobehavioral
abnormalities,
the
LOAEL
was
20
mg/
kg/
day,
the
lowest
dose
tested,
and
a
NOAEL
could
not
be
determined.

No
long­
term
systemic
toxicity
studies
for
any
exposure
route
were
identified
in
the
peerreviewed
literature.
However,
DBA
is
currently
undergoing
90­
day
subchronic
and
2­
year
chronic
bioassays
(
NTP,
2000b).
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
22
Draft,
do
not
cite
or
quote
C.
Reproductive
and
Developmental
Effects
Much
of
the
emphasis
on
reproductive
toxicity
of
brominated
acetic
acids
has
focused
on
the
potential
spermatotoxicity
of
these
compounds.
Therefore,
to
enhance
the
reader's
evaluation
of
the
following
study
descriptions,
a
short
summary
of
spermatogenesis
relevant
to
assessing
male
reproductive
toxicity
is
provided
here.
For
additional
information
the
reader
is
referred
to
Zenick
et
al.
(
1994),
from
which
the
following
summary
text
was
largely
developed.

The
development
of
mature
sperm
(
spermatogenesis)
is
a
multiple­
step
process
that
begins
within
the
seminiferous
tubules
in
the
testes
and
is
completed
with
the
movement
of
spermatids
through
the
caput,
corpus,
and
cauda
epididymis
for
further
functional
development
and
transport
to
the
vas
deferens.
The
seminiferous
tubule
is
comprised
of
spermatogenic
cells
and
support
cells
such
as
Sertoli
cells.
The
spermatogenic
cells
undergo
a
well­
defined
step­
wise
maturation
process.
As
the
cells
mature,
they
move
from
the
basal
membrane
of
the
seminiferous
tubule
until
eventual
release
from
the
supporting
Sertoli
cells
into
the
tubule
lumen.
The
release
of
the
spermatids
from
the
Sertoli
cells
is
termed
spermiation.
The
resulting
sperm
cells
are
transported
from
the
seminiferous
tubule
lumen
to
the
epididymis
where
they
undergo
further
functional
development,
including
the
acquisition
of
motility
and
reproductive
viability.

The
maturation
of
the
spermatogenic
cells
in
the
seminiferous
tubules
occurs
through
a
series
of
phases
and
the
increasingly­
mature
spermatogenic
cells
sequentially
develop
into
spermatogonia,
spermatocytes,
and
spermatids.
Each
of
these
three
major
developmental
phases
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
23
Draft,
do
not
cite
or
quote
includes
a
series
of
smaller
developmental
steps.
For
example,
for
the
rat
there
are
19
developmental
phases
or
"
steps"
for
spermatids.
Within
the
seminiferous
tubule,
spermatogenic
cells
in
various
steps
of
development
are
found
in
distinct
and
repeatable
associations.
Each
common
set
of
associations
is
called
a
Stage
of
the
seminiferous
epithelium.
For
example,
a
cross­
section
of
a
rat
seminiferous
tubule
at
Stage
VIII
would
typically
contain
PI
and
P
spermatocytes,
and
Step
8
and
19
spermatids.
Thus,
perturbations
in
normal
cell
associations
can
serve
as
an
indication
of
spermatotoxicity.
The
time
period
between
the
appearance
of
the
same
Stage
at
a
given
point
in
the
epithelium
is
called
the
cycle
length
of
the
seminiferous
epithelium.

The
stages
and
cycle
length
vary
across
species,
but
are
nearly
constant
in
the
same
species.
The
consistent
length
of
time
for
spermatogenesis
can
be
useful
for
identifying
targets
for
spermatotoxicity,
particularly
for
single­
dose
studies.
For
example,
information
about
potential
targets
of
toxicity
can
often
be
gained
by
determining
the
amount
of
time
from
the
time
of
exposure
to
a
toxicant
to
the
appearance
of
adverse
effects
by
tracing
back
to
the
phase
of
development
that
the
affected
sperm
was
in
at
the
time
of
exposure.

Spermatogenesis
can
also
be
perturbed
through
toxicity
directed
at
cell
populations
that
aid
in
the
maturation
of
the
spermatogenic
cells.
Sertoli
cells
provide
support
functions
and
developmental
regulation
of
sperm
cells,
and
thus
can
be
an
important
target
for
toxicity.
Sertoli
cells
play
important
roles
in
endocrine
regulation
of
spermatogenesis,
provide
a
protective
semipermeable
barrier
for
the
seminiferous
tubules,
and
provide
direct
support
for
development
of
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
24
Draft,
do
not
cite
or
quote
spermatogenic
cells
through
phagocytosis.
For
example,
the
Sertoli
cells
phagocytize
a
portion
of
the
cytoplasm
and
overlying
membrane
of
the
spermatid
to
form
a
residual
body
at
spermiation.

Monobromoacetic
acid
Linder
et
al.
(
1994a)
reported
the
results
of
acute­
toxicity
and
acute­
spermatotoxicity
studies
of
MBA.
In
the
spermatotoxicity
study,
male
Sprague­
Dawley
rats
(
8/
group)
were
given
a
single
dose
of
either
0
or
100
mg/
kg
MBA
in
a
volume
of
5
mL/
kg
in
water,
and
were
sacrificed
2
or
14
days
after
dosing.
The
selected
single
dose
of
100
mg/
kg
was
an
approximate
LD
01,
and
was
chosen
to
provide
a
relatively­
high
dose
with
a
minimal
likelihood
of
mortality.
Measures
of
male
reproductive
toxicity
included
reproductive­
organ
weights,
sperm
counts,
sperm
morphology,

sperm
motility,
and
histopathological
examination
of
the
seminiferous
tubules.
No
adverse
effects
were
observed
in
the
single­
dose
study;
therefore,
a
repeated­
dosing
protocol
experiment
was
also
conducted.
Groups
of
eight
rats
were
given
daily
doses
of
0
or
25
mg/
kg/
day
MBA
in
water
for
14
days,
and
were
sacrificed
24
hours
after
the
last
dose.
MBA
also
failed
to
induce
any
spermatotoxicity
in
this
repeated­
dosing
study.

In
a
published
abstract,
Randall
et
al.
(
1991)
reported
on
the
reproductive
and
developmental
toxicity
of
MBA.
Pregnant
Long­
Evans
rats
were
given
oral
gavage
doses
of
0,

25,
50,
or
100
mg/
kg/
day
MBA
in
distilled
water
on
gestation
days
6­
15.
In
the
high­
dose
group,

maternal
weight­
gain
was
reduced
and
one
dam
died.
No
effects
on
reproduction
were
observed.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
25
Draft,
do
not
cite
or
quote
Several
developmental
effects
were
noted
in
the
high­
dose
group,
including
decreased
size
of
live
fetuses
(
the
affected
measure
of
size
was
not
provided
in
the
study
summary)
and
increased
incidence
of
soft­
tissue
malformations,
most
of
which
were
cardiovascular
and
craniofacial.
Based
on
the
limited
data
provided
in
the
abstract,
the
LOAEL
for
both
maternal
and
developmental
effects
is
100
mg/
kg/
day
and
the
NOAEL
is
50
mg/
kg/
day.

No
reproductive
or
developmental
toxicity
for
MBA
was
identified
following
dosing
by
the
inhalation
or
dermal
routes.

Bromochloroacetic
acid
NTP
(
1998)
reported
the
results
of
a
short­
term
reproductive
and
developmental
toxicityscreening
protocol
for
BCA.
Details
of
the
protocol
for
this
study
are
provided
in
Section
V.
A.

Briefly,
male
and
female
Sprague­
Dawley
rats
were
administered
0,
60,
200,
or
600
ppm
BCA
in
their
drinking
water
for
various
periods
during
a
35­
day
study
period.
The
rats
were
divided
into
two
groups
of
males
and
three
groups
of
females.
Group
A
males
(
10/
group)
were
exposed
on
study
days
6­
35.
Group
B
males
were
exposed
on
study
days
6­
31
to
0,
60,
or
200
ppm
(
5/
group)

or
to
600
ppm
(
8/
group),
and
were
subsequently
treated
with
BrdU
for
3
days
prior
to
necropsy
to
evaluate
cell
proliferation.
The
study
authors
reported
that
the
estimated
average
doses
for
males
were
0,
5,
15,
and
39
mg/
kg/
day.
Male
rats
were
evaluated
for
clinical
pathology,
organ
weights,
sperm
analysis
(
group
A
only),
and
histopathology.
No
consistent
treatment­
related
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
26
Draft,
do
not
cite
or
quote
effects
on
epididymal
sperm
measures,
spermatid
head
counts,
sperm
morphology,
or
sperm
motility
were
observed
at
necropsy.

Among
females,
Group
A
(
10/
dose
group)
rats
were
treated
with
BCA
on
study
days
1­
34
and
cohabitated
with
treated
males
on
study
days
13­
18.
Group
B
(
13/
dose
group)
females
were
cohabitated
with
treated
males
on
study
days
1­
5
and
exposed
on
GD
6
through
parturition.

Group
C
females
(
5/
group
at
0,
60,
200
ppm,
and
8
animals
at
600
ppm)
were
exposed
to
a
dosing
regime
similar
to
that
of
Group
A,
but
were
removed
from
treatment
on
study
day
30
and
subsequently
administered
BrdU
to
assess
target­
tissue
cell
proliferation.
Thus,
the
treatment
protocol
for
Group
A
resulted
in
exposure
for
12
days
prior
to
mating
and
from
GD
1­
16
or
1­
21,

depending
on
the
number
of
days
of
cohabitation
required
for
mating.
The
treatment
protocol
for
Group
C
females
resulted
in
12
days
of
premating
exposure
and
exposure
beginning
on
GD1
and
continuing
through
GD
12­
16,
depending
on
the
number
of
days
required
for
mating.
Both
groups
were
evaluated
for
indices
of
mating
and
fertility
and
number
of
corpora
lutea,
live
and
dead
fetuses,
and
implantation
sites.
The
study
authors
reported
that
the
estimated
average
daily
doses
resulting
for
both
groups
were
0,
6,
19,
and
50
mg/
kg/
day.
No
effects
were
observed
on
the
mating
index
(
number
of
females
with
vaginal
sperm
/
number
of
cohabitating
pairs),
pregnancy
index
(
number
of
fertile
pairs
/
number
of
cohabitating
pairs)
or
fertility
index
(
number
of
fertile
pairs
/
number
of
females
with
vaginal
sperm).
Due
to
the
limited
number
of
pregnancies
evaluated
and
the
similar
dosing
protocols,
reproductive­
outcome
data
were
pooled
for
Groups
A
and
C
females.
Analysis
of
the
combined
results
revealed
statistically
significant
decreases
of
up
to
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
27
Draft,
do
not
cite
or
quote
70%
in
the
number
of
live
fetuses
per
litter
and
up
to
75%
in
total
implants
per
litter,
as
compared
with
controls.
Pre­
implantation
losses
increased
up
to
249%
of
controls
in
the
combined
highdose
group,
but
this
result
was
not
statistically
significant.
A
summary
of
selected
endpoints
for
the
combined
Group
A
and
C
female
data
is
provided
in
Table
V­
4.

Statistically
significant
treatment­
related
effects
by
individual
groups
included
a
16%

decrease
in
total
implants
per
litter
in
the
600
ppm
Group
A
females
and
a
50%
decrease
in
number
of
live
fetuses
per
litter
in
600
ppm
Group
C
females.
A
number
of
other
outcomes
for
either
Group
A
or
C
were
reported
to
be
adversely
altered
by
BCA
treatment
but
did
not
differ
statistically
from
controls:
(
1)
post­
implantation
losses
were
increased
in
the
600
ppm
Group
C
females;
(
2)
pre­
implantation
losses
were
increased
in
the
600
ppm
Group
A
females
and
all
dosed
groups
in
Group
C;
(
3)
an
increase
in
total
resorptions
was
observed
in
the
600
ppm
Group
C
females;
and
(
4)
decreased
total
implants
per
litter
occurred
in
all
dosed
groups
in
Group
C.
The
study
author
noted
that
the
reason
that
many
of
these
adverse
outcomes
lacked
statistical
significance
may
have
been
due
to
the
small
number
of
pregnancies
(
N
=
2
to
5)
per
treatment
group
evaluated
in
this
screening
protocol.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
28
Draft,
do
not
cite
or
quote
Table
V­
4.
Reproductive
and
Developmental
Toxicity
of
BCA
following
Peri­
conception
Exposure
(
Combined
Data
for
Female
Groups
A
and
C)
a
Estimated
Dose
(
mg/
kg/
day)

Parameter
0
6
19
50
Live
fetuses
per
litter
(%
of
controls)
14.9
±
1.05b
(
100%)
12.2
±
1.36
(
82%)
13.2
±
0.63
(
89%)
10.5
±
1.14*
(
70%)

Total
implants
per
litterc
(%
of
controls)
16.4
±
1.18
(
100%)
13.7
±
1.41
(
84%)
14.6
±
0.73
(
89%)
12.3
±
1.29*
(
75%)

%
Pre­
implantation
lossd
(%
of
controls)
12.92
±
6.24
(
100%)
18.52
±
6.81
(
143%)
15.27
±
4.18
(
118%)
32.17
±
7.48
(
249%)

%
Post­
implantation
losse
(%
of
controls)
8.67
±
1.62
(
100%)
11.50
±
3.10
(
133%)
8.65
±
3.16
(
100%)
12.68
±
3.88
(
146%)

Total
resorptionsf
(%
of
controls)
1.5
±
0.27
(
100%)
1.5
±
0.41
(
100%)
1.4
±
0.54
(
93%)
1.8
±
0.62
(
120%)

Dead
fetuses
per
litter
(
Number
of
pregnant
females)
0.0
±
0.00
(
10)
0.0
±
0.00
(
11)
0.0
±
0.00
(
11)
0.0
±
0.00
(
12)

Notes:

a.
Adapted
from
NTP,
1998.

b.
Mean
±
standard
error.

c.
Total
implants
=
number
of
viable
fetuses
+
early
resorptions
+
late
resorptions
+
dead
fetuses
d.
%
Pre­
implantation
loss
=
[(
corpora
lutea
­
total
implants)
/
corpora
lutea]
x
100
e.
%
Post­
implantation
loss
=
[(
resorptions
+
dead
fetuses)
/
total
implants]
x
100
f.
Total
resorptions
=
early
resorptions
+
late
resorptions
*
Statistical
significance:
p<
0.05
Group
B
females
(
cohabitation
with
males
on
study
days
1
to
5
and
exposure
on
GD
6
to
parturition)
were
assessed
for
maternal
body
weight;
feed
and
water
consumption;
number
of
uterine
implantations;
number,
weight,
and
anogenital
distance
of
pups;
and
evaluation
of
fetal
heart
and
brain
for
soft­
tissue
malformations.
The
study
authors
reported
that
the
estimated
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
29
Draft,
do
not
cite
or
quote
average
daily
doses
of
BCA
were
0,
10,
25,
or
61
mg/
kg/
day
for
this
group.
The
only
observed
effect
was
an
increase
in
post­
implantation
losses
in
all
groups,
which
decreased
with
increasing
dose,
although
all
losses
were
elevated
relative
to
controls
(
303%,
190%,
and
184%
of
the
control
value
in
the
60,
200,
and
600
ppm
groups,
respectively).
In
addition,
total
resorptions
were
increased
to
200%,
137%,
and
137%
of
controls
in
the
60,
200,
and
600
ppm
groups,

respectively.
None
of
the
effects
were
statistically
different
from
controls,
and
the
negative
doseresponse
makes
it
difficult
to
assess
the
biological
significance
of
the
findings.
No
treatmentrelated
effects
were
observed
in
soft­
tissue
examination
(
heart
and
brain)
of
the
fetuses.

Evaluation
of
the
total
data
set
of
both
significant
and
non­
significant
effects
suggested
to
the
authors
that
BCA
adversely
affected
the
ability
of
females
to
conceive
and
carry
a
full
litter
to
term.
The
effects
of
BCA
appear
to
be
particularly
relevant
for
early
gestation,
as
demonstrated
by
significantly
increased
pre­
implantation
losses
and
decreased
total
implants
per
litter,
and
nonsignificant
but
elevated
post­
implantation
losses
and
increased
number
of
resorptions.

Determination
of
a
LOAEL
and
NOAEL
for
this
study
is
undermined
by
the
small
sample
sizes
used
in
the
screening
protocol
and
the
low
number
of
pregnancies
per
dose
group.
Nonetheless,
a
number
of
reproductive
and
development
effects
of
significant
severity
were
reported
in
all
dose
groups.
Based
on
biologically­
relevant
changes
that
were
statistically
different
from
control
values,
the
LOAEL
for
reproductive
and
developmental
effects
(
reduced
implants
per
litter
and
live
fetuses
per
litter)
was
50
mg/
kg/
day
(
high­
dose
group)
and
the
NOAEL
was
19
mg/
kg/
day
(
mid­
dose
group).
It
should
be
noted,
however,
that
the
LOAEL
and
NOAEL
might
have
been
Drinking
Water
Criteria
Document
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OST/
HECD
V­
30
Draft,
do
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cite
or
quote
significantly
lower
if
the
statistical
power
of
these
experiments
had
been
increased
by
the
use
of
a
larger
sample
size.
As
discussed
in
Section
V.
A.,
the
high
dose
was
a
marginal
LOAEL,
and
19
mg/
kg/
day
was
a
NOAEL
for
maternal
toxicity.

The
effects
of
BCA
on
reproduction
in
male
mice
have
also
been
evaluated
following
oral
gavage
dosing.
Luft
et
al.
(
2000)
reported
in
an
abstract
on
a
study
in
which
male
C57BL/
6
mice
(
12
mice/
group)
were
administered
daily
gavage
doses
of
0,
8,
24,
72,
or
216
mg/
kg
BCA
for
14
days.
After
14
days,
5
mice/
group
were
necropsied
for
histopathological
examination
of
the
testes,
epididymis,
and
seminal
vesicles.
The
remaining
7
males
were
used
in
a
40­
day
breeding
assay
to
evaluate
the
effects
of
BCA
treatment
on
fertility.
Coital
plug­
positive
females
(
presumably
untreated)
were
replaced
daily
and
uteri
were
dissected
14
days
later;
the
numbers
of
implantations,
resorptions,
and
fetuses
were
determined.
No
effects
on
body
weight
or
reproductive­
organ
weights
were
observed
for
any
of
the
dose
groups.
The
results
of
histopathologic
examination
of
the
male
reproductive
tissues
were
not
reported
in
the
abstract.

However,
BCA
treatment
with
72
or
216
mg/
kg/
day
resulted
in
adverse
reproductive
performance,
but
only
during
the
first
10
days
following
treatment
(
data
not
shown).
Adverse
measures
of
reproductive
outcome
included
statistically
significant
decreases
in
both
of
the
dose
groups
for
(
1)
mean
number
of
litters
per
male
(
1.1
for
both
dose
groups
compared
to
3
for
controls);
(
2)
percentage
of
litters
per
mated
female
as
measured
by
the
percent
of
plug­
positive
females
that
became
pregnant
(
36%
and
30%
in
the
72
and
216
mg/
kg/
day
groups,
respectively,

as
compared
to
68%
for
controls);
and
(
3)
total
number
of
fetuses
per
male
(
10
and
9
in
the
72
Drinking
Water
Criteria
Document
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OW/
OST/
HECD
V­
31
Draft,
do
not
cite
or
quote
and
215
mg/
kg/
day
groups
as
compared
with
27
for
controls).
There
was
no
difference
in
the
number
of
coital
plugs,
suggesting
that
treatment
did
not
result
in
adverse
behavioral
effects
on
mating.
The
number
of
fetuses
per
litter,
number
of
resorptions,
and
number
of
terata
were
also
unaltered,
indicating
that,
under
the
conditions
of
this
study,
adverse
reproductive
effects
in
male
mice
did
not
induce
developmental
toxicity.
This
study
appears
to
have
identified
a
LOAEL
for
decreased
male
fertility
of
72
mg/
kg/
day
and
a
NOAEL
of
24
mg/
kg/
day,
but
a
definitive
conclusion
would
require
a
review
of
the
full
study.

Klinefelter
et
al.
(
2002a)
administered
BCA
in
a
dose
range
finding
study
(
dissolved
in
deionized
water
and
pH­
adjusted
to
6.5)
by
gavage
to
adult
male
Sprague­
Dawley
rats
(
12/
dose)

at
doses
of
0,
24,
72,
or
216
mg/
kg/
day
for
14
days.
The
doses
were
selected
to
represent
the
BCA
molar
equivalents
of
0,
30,
90
and
270
mg/
kg/
day
DBA,
previously
tested
in
the
same
laboratory
(
Linder
et
al.,
1994).
Endpoints
assessed
included
body
weight;
testes,
epididymes,
and
seminal
vesicle
weights;
ex
vivo
assessment
of
testosterone
production;
and
serum
levels
of
testosterone,
luteinizing
hormone
(
LH),
follicle­
stimulating
hormone
(
FSH),
and
prolactin.
Sperm
motility,
sperm
morphology
(
cauda
and
caput),
and
sperm
counts
(
testicular
sperm
head
count
and
epididymal
sperm
counts)
were
also
evaluated.
Testis
sections
were
examined
by
light
microscopy
for
delayed
spermiation,
formation
of
atypical
residual
bodies,
and
germ
cell
depletion.

Body
weight
was
significantly
decreased
in
the
highest
dose
group.
Testis,
epididymis,
and
seminal
vesicle
weights
were
unaffected
by
BCA
treatment.
While
spermatid
numbers
were
not
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
32
Draft,
do
not
cite
or
quote
altered
by
BCA
exposure,
a
significant
dose­
related
decline
in
epididymal
sperm
reserves
was
observed
at
72
and
216
mg/
kg/
day,
with
the
effect
on
cauda
epididymal
sperm
being
more
severe
than
on
caput
epididymal
sperm.
Dose­
related
decreases
in
serum
LH,
FSH,
and
prolactin
were
noted
in
all
dosed
groups,
with
statistical
significance
occurring
in
the
two
highest
dose
groups.

No
effects
on
testis
sperm
production
or
serum
testosterone
were
observed.

The
percentage
of
motile
and
progressively
motile
cauda
sperm
decreased
in
a
doserelated
fashion,
achieving
significance
in
the
two
highest
dose
groups.
Sperm
motion
parameters
(
i.
e.,
velocity
and
linearity)
were
similarly
affected.
A
dose­
dependent
reduction
in
the
percentage
of
morphologically
normal
cauda
and
caput
epididymal
sperm
was
also
observed.
For
cauda
sperm,
the
percent
normal
sperm
decreased
to
33%
in
the
216
mg/
kg/
day
group,
as
compared
with
98.3%
in
controls.
A
similar
decrease
occurred
in
caput
sperm,
with
the
percent
normal
sperm
being
31.2%
in
the
216
mg/
kg/
day
dose
group,
as
compared
with
94.8%
in
controls.
Caput
epididymal
sperm
abnormalities
were
characterized
by
an
increased
number
of
sperm
with
misshapen
heads
or
tail
defects,
whereas
cauda
sperm
abnormalities
consisted
mainly
of
an
increased
number
of
isolated
heads.
Histological
evaluation
of
the
testis
showed
a
dose­
related
increase
(
statistically
significant
in
the
two
highest
dose
groups)
in
the
number
of
Step
19
spermatids
retained
in
Stage
X
and
XI
of
the
spermatogenic
cycle.
Other
findings
included
a
doserelated
increase
in
the
number
and
size
of
atypical
residual
bodies
in
Stages
X
and
XI
(
not
quantified)
and
a
shift
in
localization
of
these
bodies,
from
basal
migration
to
luminal
release,
with
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
33
Draft,
do
not
cite
or
quote
increasing
BCA
dose.
According
to
the
study
authors,
the
LOAEL
for
altered
spermiation
in
this
study
was
24
mg/
kg/
day,
the
lowest
dose
tested,
and
a
NOAEL
could
not
be
determined.

In
a
subsequent
definitive
study
by
the
same
authors
(
Klinefelter
et
al.,
2002b),
adult
male
Sprague­
Dawley
rats
(
10/
dose)
were
administered
14
daily
gavage
doses
of
BCA
(
dissolved
in
deionized
water
and
pH­
adjusted)
of
0,
8,
24,
or
72
mg/
kg/
day.
End
points
evaluated
were
the
same
as
those
assessed
in
the
previous
study.
Additionally,
sperm
protein
was
extracted
and
analyzed,
and
a
fertility
assessment
was
conducted
via
in
utero
insemination
of
untreated
females
with
sperm
from
treated
males.

For
the
fertility
assessment,
the
estrus
cyclicity
of
a
cohort
of
females
was
synchronized
by
administering
a
subcutaneous
injection
of
an
luteinizing
hormone
releasing
hormone
(
LHRH)

agonist
at
115
hours
prior
to
insemination.
At
the
beginning
of
the
dark
cycle
following
proestrus,

each
female
was
paired
with
a
sexually
experienced
vasectomized
male
for
30
minutes.
Receptive
females
(
as
indicated
by
the
presence
of
a
copulatory
plug)
were
subsequently
anaesthetized,
and
epididymal
sperm
from
treated
males
were
injected
into
each
uterine
horn
at
an
amount
(
5
x
106)

that
results
in
approximately
75%
fertility
in
control
animals.
The
sperm
from
a
single
male
was
used
to
inseminate
a
single
female.
Inseminated
females
were
sacrificed
9
days
following
treatment,
and
implanted
embryos
and
corpora
lutea
of
pregnancy
were
counted.
Male
fertility
was
expressed
as
a
percentage
equivalent
to
the
number
of
implants/
corpora
lutea
x
100.

No
treatment­
related
changes
in
body
weight,
testes
weight,
and
the
weight
of
the
seminal
vesicles
were
observed.
However,
in
contrast
with
the
previous
study
conducted
by
the
same
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
34
Draft,
do
not
cite
or
quote
authors,
epididymal
weights
were
reduced
at
72
mg/
kg/
day,
and
there
were
no
differences
between
treated
and
control
groups
in
any
of
the
hormonal
measurements.
Sperm
motion
parameters
were
consistently
altered
by
BCA
exposure.
Although
the
percentage
of
motile
sperm
was
only
decreased
in
the
high­
dose
group
(
72
mg/
kg/
day),
progressive
sperm
motility
was
decreased
at
all
doses
tested.
Altered
sperm
morphology
was
only
observed
at
72
mg/
kg/
day;

abnormalities
in
both
cauda
and
caput
sperm
were
similar
to
those
observed
in
the
earlier
study,

with
the
cauda
sperm
showing
increased
incidences
of
sperm
with
tail
defects
and
the
caput
sperm
showing
increased
incidences
of
sperm
with
isolated
heads.
In
utero
insemination
of
untreated
females
with
the
cauda
epididymal
sperm
from
treated
males
showed
a
significant
reduction
in
fertility
at
all
doses,
but
no
dose­
response.
Fertility
rates
in
the
8,
24,
and
72
mg/
kg/
day
groups
were
33%,
44%,
and
37%,
respectively,
as
compared
with
75%
in
control
animals.
The
LOAEL
for
this
study
was
8
mg/
kg/
day,
the
lowest
dose
tested,
and
a
NOAEL
could
not
be
determined.

In
the
sperm
protein
extraction
phase
of
the
study,
two­
dimensional
evaluation
of
120
proteins
showed
significant
reductions
in
two
proteins,
SP22
and
SP9.
The
shape
of
the
doseresponse
curve
for
SP22
paralleled
the
reduction
in
fertility,
whereas
that
for
SP9
did
not.
The
Pearson
correlation
coefficient
was
0.53
(
p
<
0.001)
for
SP22
and
fertility,
and
0.23
for
SP9
and
fertility.
When
the
data
were
fitted
to
a
non­
linear
threshold
response
model,
the
resulting
correlation
coefficient
(
r2)
for
SP22
and
fertility
was
0.843.
The
study
authors
concluded
that
BCA,
like
DBA,
is
capable
of
perturbing
spermatogenesis
and
fertility,
and
that
SP22
appears
to
be
useful
as
a
sperm
biomarker
of
fertility.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
35
Draft,
do
not
cite
or
quote
No
reproductive­
or
developmental­
toxicity
studies
for
BCA
were
identified
following
dosing
by
the
inhalation
or
dermal
routes.

Dibromoacetic
acid
There
has
been
considerable
interest
in
the
male
reproductive
effects
of
DBA,
in
part
because
its
chlorinated
analog,
dichloroacetic
acid
(
DCA),
is
known
to
be
a
male
reproductive
toxicant.
Linder
et
al.
(
1994a)
reported
the
results
of
acute­
toxicity
and
acute­
spermatotoxicity
studies
of
DBA.
A
single­
dose
protocol
was
used
to
identify
stages
of
spermatogenesis
that
might
be
impacted
by
DBA.
In
the
spermatotoxicity
study,
male
Sprague­
Dawley
rats
(
8/
group)
were
administered
a
single
gavage
dose
of
0
or
1250
mg/
kg
DBA
and
sacrificed
2,
14,
or
28
days
after
dosing.
The
approximate
LD
01
dose
of
1250
mg/
kg
was
selected
to
provide
a
relatively­
high
dose
with
minimal
likelihood
of
mortality.
The
study
duration
was
extended
to
28
days
because
of
evidence
from
the
acute­
toxicity
study
that
effects
on
epididymal
sperm
could
peak
more
than
14
days
after
dosing.
Reproductive­
organ
weights
and
sperm­
quality
parameters
were
measured,
and
a
histopathologic
examination
was
performed.
Only
marginal
reproductive­
organ­
weight
changes
were
induced
by
DBA.
Epididymis
weights
on
Days
2
and
28
were
decreased
to
93%
and
83%
of
control
values
(
p<
0.05),
respectively,
but
were
not
different
from
controls
on
Day
14.
Testes
weights
were
decreased
to
93%
of
controls
on
Day
28
(
p<
0.05).
Prostate
weights
were
significantly
increased
to
109%
of
controls
on
Day
2
(
p<
0.05).
DBA
treatment
did
not
affect
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
36
Draft,
do
not
cite
or
quote
body
weights,
suggesting
that
reproductive­
organ­
weight
changes
were
not
secondary
to
general
toxicity.
Serum­
testosterone
levels
fell
to
17%
of
control
values
2
days
after
a
single
dose
of
1250
mg/
kg,
but
returned
to
control
levels
by
Day
14.

Several
measures
of
spermatotoxicity
were
reported
in
this
study.
Caput­
sperm
count
was
significantly
reduced
on
Day
2
to
85%
of
controls,
but
was
not
affected
on
Days
14
or
28.

Caudasperm
count
was
decreased
to
54%
and
44%
of
controls
on
study
Days
14
and
28,
respectively.

Testicular­
sperm
head
count
was
not
affected,
suggesting
that
DBA
was
not
inhibiting
overall
sperm
production.
Sperm
morphology
was
also
seriously
affected
by
exposure
to
DBA.
The
percent
of
sperm
having
flagellar
defects
and
atypical
heads
was
significantly
increased
in
caput
sperm
on
Day
28,
with
16%
showing
abnormal
morphology.
In
the
control
group,
about
5%
of
sperm
were
estimated
to
be
abnormal,
based
on
direct
inspection
of
the
data
presented
in
a
figure
in
the
paper.
Cauda
sperm
showed
a
dramatic
increase
in
flagellar
defects
on
Day
14
(
p<
0.05),
but
not
on
Day
28.
Significant
increases
in
sperm
with
atypical
heads
and
with
both
atypical
head
and
flagellar
defects
were
increased
on
Day
28
(
64%
of
sperm
displayed
abnormal
morphology).

According
to
the
study
authors,
the
appearance
of
different
morphological
changes
(
flagellar
versus
acrosomal)
on
Days
14
and
28
indicated
that
the
epididymal
sperm
underwent
two
sequential
morphological
changes
as
a
result
of
DBA
exposure.
Several
measures
of
sperm
motility
were
significantly
reduced
at
Day
14,
including
percent
motile
(
38%
of
controls),
percent
progressive
motility
(
32%
of
controls),
straight­
line
velocity
(
73%
of
controls),
and
curvilinear
velocity
(
82%
of
controls).
Only
the
first
two
measures
were
significantly
reduced
at
Day
28
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
37
Draft,
do
not
cite
or
quote
(
51%
and
41%
of
controls,
respectively).
The
study
authors
suggested
that
these
decreases
were
related
to
the
flagellar
defects.
Thus,
treatment
with
a
single
dose
of
1250
mg/
kg
of
DBA
resulted
in
significantly
adverse
effects
on
sperm
count,
morphology,
and
motility.

Histopathology
examination
revealed
altered
spermiation
at
all
three
time­
points
examined
(
Days
2,
14,
and
28).
On
all
three
days,
Step
19
spermatids
were
retained
beyond
their
normal
release
in
Stage
VIII
of
the
seminiferous
epithelium
cycle.
Other
abnormal
histological
signs
included
the
presence
of
remnants
of
residual
bodies
in
Stages
X
and
XI,
and
the
presence
of
anucleate
cytoplasmic
debris
in
the
lumen
of
the
epididymal
duct
and
in
the
caput
epididymis
on
Day
2.
On
Day
14,
debris
from
the
testes
was
evident
in
the
epididymis,
much
of
which
resembled
residual
bodies.
On
Days
14
and
28,
abnormal
late
spermatids
were
observed
in
Stages
I­
VIII.

Similar
histological
changes
were
observed
on
both
days,
although
the
changes
were
characterized
by
the
authors
as
less
severe
on
Day
28.
Varying
amounts
of
cytoplasmic
debris
were
also
observed
in
the
epididymis
on
Day
28.

These
results
show
that
a
single
high
dose
of
DBA
is
spermatotoxic
in
the
rat.
The
target
cells
for
adverse
effects
of
DBA
were
not
conclusively
identified,
although
the
authors
described
several
aspects
of
sperm
maturation
that
might
be
impacted,
based
on
consideration
of
the
normal
transit
times
(
assuming
that
the
kinetics
of
sperm
development
were
not
affected),
and
on
consideration
of
the
timing
of
the
observed
effects.
The
study
authors
noted
that
flagellar
degeneration
was
observed
in
the
cauda
epididymis
on
Day
14,
but
not
in
the
caput
epididymis.

The
flagellar
changes
might
be
due
to
a
DBA
effect
during
transit
through
the
epididymis.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
38
Draft,
do
not
cite
or
quote
Alternatively,
as
the
normal
transit
time
of
sperm
through
the
caput
epididymis
is
3
days
in
the
rat,

and
accounting
for
the
remaining
11
days
since
dosing,
the
authors
suggested
that
late
spermatids
before
or
during
spermiation
might
have
been
affected.
On
day
28,
both
altered
sperm
heads
and
flagellar
degeneration
were
observed
in
both
the
caput
and
cauda
sperm.
As
spermatids
that
were
in
Step
11
to
15
on
Day
2
of
exposure
would
have
comprised
the
majority
of
caput
sperm
on
Day
14
(
when
only
a
minimal
effect
on
head
development
was
seen),
the
authors
suggested
that
the
abnormal
head
development
may
have
resulted
from
an
effect
of
DBA
on
Step
10
or
earlier
spermatids.
Alternatively,
they
noted
that
the
same
effects
would
have
been
seen
if
the
effect
of
DBA
was
on
later
steps
but
the
action
of
DBA
was
delayed
for
several
days
following
dosing.
As
the
retention
of
Step
19
spermatids
is
an
effect
observed
following
treatments
that
alter
hormone
status,
the
observation
that
DBA
reduced
circulating­
testosterone
levels
is
consistent
with
the
effects
on
Step
19
spermatids
noted
in
the
study.
According
to
the
authors,
another
potential
target
for
DBA
might
be
Sertoli
cells,
since
the
presence
of
testicular
debris
might
suggest
disruption
of
the
endocytic
activity
of
these
cells.
While
DBA
treatment
adversely
affected
sperm
quality,
it
did
not
appear
to
inhibit
sperm
production,
based
on
histological
analysis
of
the
testes
and
the
absence
of
an
effect
on
testicular
sperm­
head
counts.

Linder
et
al.
(
1994b)
studied
the
spermatotoxicity
of
DBA
following
14
daily
exposures.

The
effects
on
selected
endpoints
are
presented
in
Table
V­
5.
Male
Sprague­
Dawley
rats
(
8/
dose
group),
approximately
four
months
old,
were
given
daily
gavage
DBA
doses
of
0,
10,
30,
90,
or
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
39
Draft,
do
not
cite
or
quote
270
mg/
kg/
day,
and
sacrificed
immediately
after
the
last
dose.
The
dose
vehicle
was
distilled
water
and
the
dose
volume
was
5
mL/
kg,
adjusted
weekly
for
body
weight.
No
effects
on
body
weight
or
serum­
testosterone
levels
were
noted
at
any
dose
level.
Several
parameters
were
affected
(
p<
0.05),
primarily
at
the
highest
dose
level
of
270
mg/
kg.
These
included
mildly
reduced
testis
(
93%
of
controls)
and
epididymis
weights
(
86%
of
controls).
Absolute
and
relative
(
to
testis
weight)
testicular
sperm­
head
counts
were
also
repressed
to
81%
and
88%
of
control
values,

respectively.
Dose­
dependent
effects
on
various
measures
of
sperm
motility
were
also
observed,

with
statistically
significant
decreases
observed
at
the
highest
dose.
Caput­
sperm
counts
were
reduced
significantly
in
a
dose­
dependent
fashion
beginning
at
the
low
dose
of
10
mg/
kg/
day.
The
percent
of
morphologically­
normal
sperm
was
statistically
decreased
(
79%
of
controls)
only
in
the
high­
dose
group,
with
atypical
heads
observed
more
frequently
than
degenerating
flagella
and
a
notable
increase
detected
in
fused
sperm.
Cauda­
sperm
count
was
significantly
reduced
to
76%

and
30%
of
control
values
in
the
90
and
270
mg/
kg/
day
dose
groups,
respectively.
The
percent
of
morphologically­
normal
sperm
was
decreased
to
86%
and
32%
of
controls
in
the
same
groups.

Morphological
changes
in
cauda
sperm
were
mainly
related
to
degenerative
changes
of
the
flagella.
Percent
motile
sperm
and
percent
progressive
motility
were
reduced
to
less
than
10%
of
control
values
at
270
mg/
kg/
day.
Straight­
line
velocity
and
linearity
were
also
reduced
at
the
highest
dose.
Curvilinear
velocity
was
significantly
affected
only
at
90
mg/
kg/
day.

Histopathological
evidence
of
altered
spermiation
was
noted
beginning
at
10
mg/
kg/
day.

Histopathological
findings
included
retention
of
Step
19
spermatids
in
Stages
IX
to
XII
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
40
Draft,
do
not
cite
or
quote
atypical
acrosomal
development
of
Step
15
spermatids
at
10
mg/
kg/
day.
The
severity
of
these
effects
increased
with
increasing
dose.
The
presence
of
atypical
structures
resembling
residual
bodies
in
the
testis
and
caput
epididymis
was
observed
at
the
two
highest
doses.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
41
Draft,
do
not
cite
or
quote
Table
V­
5.
Sperm
Quality
Parameters
in
Rats
Given
14
Daily
Doses
of
DBAa
Dose
(
mg/
kg/
day)

Parameter
0
10
30
90
270
Caput
sperm
106
131
±
14
b
112
±
13*
118
±
10*
108
±
11*
100
±
10*

Cauda
sperm
106
264
±
21
250
±
70
243
±
34
200
±
59*
78
±
15*

Caput
sperm
(%
morphologically
normal)
95
±
4
90
±
11
96
±
1
93
±
3
75
±
12*

Cauda
sperm
(%
morphologically
normal)
98
±
1
90
±
15
96
±
2
84
±
10*
31
±
17*

Percent
motile
78
±
7
74
±
14
83
±
6
66
±
17
6
±
7*

Progressive
motility
(%)
67
±
4
61
±
16
68
±
12
56
±
16
4
±
5*

Straight­
line
velocity
(
µ
m/
sec)
75
±
12
68
±
17
66
±
16
63
±
15
44
±
22*

Curvilinear
velocity
(
µ
m/
sec)
154
±
7
145
±
10
136
±
16
128
±
25*
141
±
42
Linearity
52
±
7
51
±
9
49
±
8
50
±
5
30
±
10**

Notes:

a.
Adapted
from
Linder
et
al.,
1994b
b.
Mean
±
standard
deviation.

*
Statistical
significance:
p<
0.05
**
Statistical
significance:
p<
0.01
In
summary,
a
variety
of
male
reproductive­
tract
toxicity
parameters
were
affected
by
DBA
in
this
study.
Adverse
spermatogenic
effects
were
noted
beginning
at
the
lowest
dose
of
10
mg/
kg/
day
and
generally
increased
in
severity
with
increasing
dose.
Mildly
decreased
caputsperm
count
was
also
observed
at
this
dose,
but
the
effect
was
not
clearly
dose­
dependent,
with
similar
decrements
(
ranging
from
76%
to
90%
of
control
values)
observed
at
all
doses.
Decreased
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
42
Draft,
do
not
cite
or
quote
cauda­
sperm
count,
frequency
of
abnormal
cauda­
sperm
morphology,
and
changes
in
sperm
motility
were
statistically
significant
only
at
the
two
highest
doses,
90
and
270
mg/
kg/
day.

Noticeable
histopathological
changes
(
delayed
release
of
Step
19
spermatids
and
atypical
Step
15
spermatid
acrosomal
development)
began
at
the
low
dose.
The
LOAEL
for
this
study
is
the
lowest
dose
tested,
10
mg/
kg/
day,
based
on
histopathological
changes
in
the
male
reproductive
tract,
and
a
NOAEL
could
not
be
determined.

Linder
et
al.
(
1995)
studied
the
longer­
term
(
up
to
79
days)
effects
of
DBA
in
male
rats
on
both
reproductive
competence
(
summarized
in
Table
V­
6
and
Table
V­
7)
and
on
sperm
quality
(
summarized
in
Table
V­
8
and
Table
V­
9).
The
highest
dose
of
250
mg/
kg/
day
was
selected
based
on
the
expectation
that
it
would
produce
substantial
spermatotoxicity
and
permit
the
investigation
of
the
time
course
of
DBA
effects
on
fertility
and
reproductive
competence.
Lower
doses
of
2,

10,
or
50
mg/
kg/
day
were
selected
to
obtain
dose­
response
data.
Selected
doses
were
based
on
the
results
of
previous
short­
term
studies
(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b),
which
suggested
that
both
no­
efect
and
significant­
effect
dose
levels
would
fall
within
this
dose
range.

Daily
doses
of
custom­
synthesized,
high­
purity
DBA
in
a
distilled­
water
vehicle
were
given
by
gavage
to
105­
day­
old
male
Sprague­
Dawley
rats
whose
reproductive
competence
had
been
proven.
There
were
essentially
two
experimental
protocols
employed,
each
of
which
will
be
reviewed
separately
here.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
43
Draft,
do
not
cite
or
quote
Table
V­
6.
Reproductive
Outcomes
in
Rats
Following
Oral
Dosing
with
DBAa
Mated
Days
Dose
mg/
kg
Male
Female
Cop
Pairsb
Copulatory
Plugsc
Sperm
positive
females
Fertile
males
No.
litters
Implants
Fetuses
8­
14
0
250
10
10
10
10
9
3*
e
3.3
±
1.3d
1.3
±
1.5
9
3*
9
2*
9
2*
14.1
±
1.6
5.5
±
6.4*
13.0
±
1.3
5.0
±
5.7*

15­
21
0
250
10
10
10
10
10
5*
3.6
±
1.1
1.4
±
0.9*
10
4*
10
0*
10
0*
13.7
±
2.3
­­­­
12.6
±
2.9
­­­

30­
37
0
2
10
50
250
10
10
10
10
10
10
10
10
10
10
9
9
7
7
7
3.6
±
2.1
3.1
±
1.3
2.6
±
1.4
4.3
±
2.2
1.4
±
1.1*
9
9
7
7
1*
9
8
7
7
0*
9
8
7
7
0*
15.7
±
2.5
15.3
±
3.1
16.3
±
1.7
13.0
±
5.2
­­­
13.6
±
3.2
14.3
±
3.5
14.9
±
2.0
11.9
±
5.3
­­­

65­
71
0
2
10
50
250
10
10
10
10
9
20
20
20
20
9
16
17
10*
13
6
3.4
±
1.3
3.3
±
1.3
2.6
±
1.8
2.4
±
1.8
1.3
±
0.8*
14
16
10
13
2*
9
10
7
9
0*
15
(
6)
e
14
(
4)
9
(
2)
10
(
1)*
0*
14.8
±
1.4
15.9
±
1.3
15.8
±
2.9
14.9
±
2.2
­­­
14.0
±
1.4
15.1
±
1.2
14.6
±
2.4
14.1
±
2.4
­­­

199­
213
0
250
10
9
20
18
15
15
3.5
±
2.0
2.9
±
1.4
15
14
10
3*
15
(
5)
5*
(
2)
14.9
±
1.7
15.3
±
1.8
14.4
±
1.9
14.5
±
2.2
a.
Adapted
from
Linder
et
al.,
1995.

b.
Copulatory
pairs
as
evidenced
by
the
presence
of
copulatory
plug
or
birth
of
a
litter.

c.
Per
copulating
pair
d.
Mean
±
SD.

e.
Numbers
in
parentheses
are
the
number
of
males
siring
two
litters.

*
Statistical
significance:
p<
0.05
Drinking
Water
Criteria
Document
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Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
44
Draft,
do
not
cite
or
quote
Table
V­
7.
Outcome
of
Artificial
Insemination
of
Sperm
from
Rats
Dosed
with
DBAa
Day
Dose
mg/
kg
Number
of
Inseminations
Number
of
litters
Implants
Fetuses
9
0
250
6
6
5
5
5.40
±
3.36b
6.00
±
3.32
5.20
±
3.63
6.00
±
3.32
16
0
250
6
5
(
1)
c
6
1*
7.83
±
4.79
4.00
7.67
±
4.63
4.00
31
0
2
10
50
250
6
6
6
6
1
(
5)
5
6
3
5
0
7.40
±
4.77
9.50
±
5.24
5.33
±
2.08
5.80
±
3.11
­­­
7.40
±
4.77
9.17
±
4.96
5.00
±
1.73
5.80
±
3.11
­­­

79
0
2
10
50
10
9
(
1)
10
10
7
6
8
5
7.86
±
2.91
7.17
±
4.26
8.75
±
2.43
9.00
±
2.83
7.71
±
2.69
6.83
±
4.45
8.63
±
2.67
9.00
±
2.83
Notes:

a.
Adapted
from
Linder
et
al.,
1995.

b.
Litter
means
±
SD.

c.
Number
in
parentheses
is
the
number
of
males
with
insufficient
sperm
for
insemination.

*
Statistical
significance,
p<
0.05
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
45
Draft,
do
not
cite
or
quote
Table
V­
8.
Reproductive
Organ
Weights
and
Sperm
Counts
in
Rats
Given
Daily
Doses
of
DBAa
Number
of
Doses
mg/
kg
Nb
Body
weight
(
g)
Testis
weight
(
g)
Epididymis
weight
(
g)
TSHCc
(
million)
TSHC
per
gram
testis
(
million)
Caput
sperm
(
million)
Cauda
sperm
(
million)
Serum
testosterone
(
ng/
ml)

2
0
250
6
6
402
±
17d
401
±
14
1.97
±
0.13
1.92
±
0.09
0.64
±
0.02
0.63
±
0.02
271
±
17
278
±
15
149
±
10
157
±
8
124
±
8
115
±
9
254
±
28
278
±
41
8.0
±
4.1
4.1
±
2.1
5
0
250
6
6
400
±
16
405
±
14
1.89
±
0.14
1.85
±
0.29
0.62
±
0.05
0.58
±
0.08
261
±
26
273
±
56
151
±
12
158
±
10
115
±
12
97
±
24
239
±
29
230
±
72
10.7
±
4.9
6.9
±
6.1
9
0
250
6
6
411
±
21
404
±
28
1.94
±
0.08
1.87
±
0.14
0.62
±
0.03
0.60
±
0.05
295
±
15
276
±
34
162
±
7
156
±
8
124
±
7
119
±
14
250
±
23
225
±
42
3.4
±
1.7
3.4
±
3.2
16
0
250
5
6
407
±
23
375
±
58
1.96
±
0.21
1.74
±
0.18
0.64
±
0.04
0.51
±
0.07*
288
±
26
263
±
27
162
±
19
162
±
11
122
±
27
92
±
21
255
±
23
75
±
32*
4.1
±
3.5
2.9
±
1.5
31
0
2
10
50
250
6
5
6
6
6
432
±
14
417
±
16
424
±
21
425
±
22
368
±
38*
1.93
±
0.12
1.93
±
0.08
1.94
±
0.15
1.90
±
0.14
1.81
±
0.05
0.64
±
0.03
0.65
±
0.02
0.64
±
0.06
0.60
±
0.02
0.51
±
0.03
265
±
24
280
±
26
264
±
35
282
±
28
283
±
25
147
±
8
155
±
10
145
±
16
160
±
11
167
±
13*
128
±
9
125
±
13
121
±
9
111
±
10*
47
±
9*
240
±
14
225
±
44
225
±
34
169
±
34*
33
±
5*
4.0
±
3.3
11.8
±
6.4
5.6
±
5.7
3.9
±
2.5
2.6
±
1.6
79
0
2
10
50
10
10
10
10
458
±
24
455
±
16
446
±
20
434
±
24*
2.01
±
0.15
1.96
±
0.12
2.02
±
0.14
1.97
±
0.11
0.68
±
0.03
0.67
±
0.03
0.69
±
0.04
0.64
±
0.05
298
±
32
270
±
25
288
±
21
289
±
24
159
±
11
148
±
11*
155
±
6
159
±
11
126
±
7
126
±
10
122
±
10
112
±
11*
240
±
40
247
±
28
242
±
34
196
±
47*
10.5
±
8.4
7.1
±
3.8
6.2
±
4.9
3.8
±
1.6
0e
0
10
522
±
22
2.02
±
0.10
0.68
±
0.06
260
±
23
139
±
6
118
±
9
249
±
45
2.7
±
1.3
42f
250
9
483
±
19*
0.99
±
0.48*
0.49
±
0.09*
50
±
106*
34
±
56*
20
±
40*
37
±
74*
2.7
±
1.5
Notes:

a.
Adapted
from
Linder
et
al.,
1995.

b.
N=
number
of
rats.

c.
TSHC
=
testicular
sperm
head
count
d.
Group
mean
±
SD.

e.
Nondosed
controls.

f.
Given
42
doses
then
allowed
to
recover
for
186
days.

*
Statistical
significance:
p<
0.05
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
46
Draft,
do
not
cite
or
quote
Table
V­
9.
Sperm
Quality
Parameters
in
Rats
Given
Daily
Doses
of
DBAa
Number
of
Doses
mg/
kg
N
b
Motile
sperm
(%)
Progressive
motility
(%)
VSLc
(
µ
m/
sec)
VCLc
(
µ
m/
sec)
Linearity
Caput
sperm
%
normal
Cauda
sperm
%
normal
2
0
250
6
6
84
±
5d
87
±
7
75
±
7
75
±
5
80
±
7
79
±
7
105
±
9
107
±
7
48
±
4
46
±
5
96
±
1
(
0)
d
96
±
2
(
0)
96
±
2
(
0)
e
97
±
1
(
0)

5
0
250
6
6
84
±
5
83
±
8
72
±
7
70
±
7
80
±
18
77
±
9
108
±
17
107
±
11
46
±
7
44
±
3
97
±
2
(
0)
90
±
4*(
3)
96
±
2
(
0)
95
±
4
(
0)

9
0
250
6
6
84
±
8
84
±
7
72
±
8
73
±
8
87
±
17
79
±
15
120
±
21
107
±
20
47
±
4
49
±
3
96
±
2
(
0)
55
±
17*
(
8)
96
±
1
(
0)
95
±
2
(
0)

16
0
250
5
6
86
±
5
13
±
18*
70
±
6
10
±
15*
63
±
24
31
±
13*
92
±
36
51
±
24*
43
±
3
32
±
10*
98
±
1
(
0)
61
±
37*
(
2)
97
±
1
(
0)
33
±
20*
(
3)

31
0
2
10
50
250e
6
5
6
6
6
78
±
3
81
±
4
72
±
20
47
±
32*
3
±
6
61
±
7
65
±
6
53
±
16
36
±
28
2
±
4*
67
±
15
74
±
10
68
±
16
55
±
28
­­­
107
±
22
113
±
16
116
±
31
91
±
36
­­­
39
±
5
42
±
4
41
±
5
35
±
8
­­­
96
±
2
(
0)
97
±
1
(
0)
96
±
6
(
0)
88
±
7*
(.
3)
1
±
2*
(
5)
96
±
1
(
0)
96
±
2
(
0)
96
±
4
(
0)
88
±
10*
(
0)
2
±
2*
(
3)

79
0
2
10
50
10
10
10
10
75
±
8
77
±
8
76
±
11
66
±
13
61
±
8
63
±
8
62
±
9
53
±
12
70
±
9
72
±
11
76
±
10
64
±
9
107
±
11
106
±
13
115
±
17
96
±
14
42
±
3
42
±
4
42
±
2
41
±
4
97
±
1
(
0)
94
±
6
(
0.02)
96
±
2
(
0)
68
±
20*
(
1)
96
±
2
(
0)
93
±
8
(
0)
94
±
4
(
0)
75
±
17*
(
0.06)

Notes:

a.
Adapted
from
Linder
et
al.,
1995.

b.
N
=
number
of
rats.

b.
VSL
is
straight
line
velocity
and
VCL
is
curvilinear
velocity.

c.
Group
means
±
SD.

d.
Number
in
parentheses
is
the
percent
of
fused
sperm.

e.
Motile
sperm
(
15%
and
1%
motile)
were
present
in
only
two
rats.

*
Statistical
significance:
p<
0.05
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
47
Draft,
do
not
cite
or
quote
In
the
first
protocol,
groups
of
10
male
rats
were
given
daily
gavage
doses
of
either
0
or
250
mg/
kg/
day
DBA.
During
study
days
8­
14,
15­
21,
and
30­
37,
the
males
were
paired
with
females
and
allowed
to
mate
by
natural
insemination.
Dosing
was
terminated
after
42
days
because
of
the
onset
of
overt
toxicity,
including
labored
breathing,
light
tremor,
difficulty
moving
the
hind
limbs,
and
severe
weight
loss.
The
animals,
however,
were
allowed
to
mate
during
recovery,
on
days
49­
56,
65­
71,
and
199­
213.
During
the
mating
period
on
study
days
8­
14,
only
3/
10
males
copulated;
only
two
males
were
fertile
and
only
two
litters
were
produced.
The
numbers
of
implants
and
fetuses
in
these
litters
were
reduced
by
more
than
50%
as
compared
to
controls.
During
the
mating
periods
on
days
15­
21
and
30­
37,
there
were
no
fertile
males
and
no
litters
were
produced,
even
though
5/
10
and
7/
10
males,
respectively,
copulated
during
these
periods.
To
distinguish
fertilization
failure
from
pre­
implantation
loss,
females
from
the
mating
period
on
days
49­
56
were
sacrificed
on
GD
1
and
examined
for
the
presence
of
fertilized
eggs.

There
were
no
fertilized
eggs.
During
the
mating
period
on
days
65­
71,
no
males
were
fertile
and
no
litters
were
produced.
During
the
mating
period
on
days
199­
213
(
after
5
months
of
recovery),

only
3/
9
males
were
fertile
and
5
litters
were
produced,
even
though
all
males
copulated.
Thus,

although
the
reproductive
performance
of
animals
in
the
250
mg/
kg/
day
improved
significantly
during
recovery,
they
never
fully
recovered.

Artificial
insemination
of
luteinizing
hormone
releasing
hormone
(
LHRH)­
synchronized
females,
with
sperm
from
treated
males
from
an
additional
group
of
6
animals,
sacrificed
on
Days
9,
16,
and
31
(
and
also
used
for
interim
necropsy),
was
conducted
to
distinguish
behavioral­
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
48
Draft,
do
not
cite
or
quote
mating
effects
from
effects
due
to
physiologic
reproductive
competence.
Five
litters
were
sired
with
Day
9
sperm;
no
significant
adverse
effects
were
observed
on
the
number
of
implants
or
fetuses.
The
absence
of
a
significant
effect
on
fertility
by
artificial
insemination
with
Day
9
sperm
suggested
to
the
study
authors
that
the
reproductive
effects
observed
from
the
Day
8­
14
mating
may
have
been
due
to
a
transient
effect
on
libido.
In
contrast,
only
one
litter
was
produced
as
a
result
of
artificial
insemination
with
Day
16
sperm,
and
no
litters
resulted
from
insemination
with
Day
31
sperm,
indicating
that
reproductive
incompetence
was
due
to
spermatotoxic
effects.

Consistent
with
these
reproductive
outcomes,
measures
of
sperm
motility
were
not
affected
until
Day
16,
but
were
severely
affected
thereafter.
The
cauda­
sperm
count
was
normal
until
Day
16,
at
which
time
it
was
reduced
to
29%
of
controls.
Caput­
sperm
counts
decreased
progressively
to
as
low
as
37%
of
normal
on
Day
31.
The
percent
of
sperm
with
normal
morphology
was
significantly
decreased
(
p<
0.05)
beginning
on
Day
5
for
caput
sperm
and
on
Day
16
for
cauda
sperm.
Only
minimal
developmental­
toxicity
data
were
provided
in
the
paper.
No
effects
on
fetal
weight
were
observed;
minimal
changes
in
the
incidence
of
malformations
were
inconsistently
observed
and
were
not
considered
to
be
treatment­
related
by
the
study
authors.

In
the
second
protocol,
groups
of
10
male
rats
were
given
daily
gavage
doses
of
0,
2,
10,

or
50
mg/
kg/
day
DBA
for
up
to
79
days.
The
only
systemic
effect
was
a
slight
decrease
in
body
weight
(
to
95%
of
controls)
that
was
apparent
by
Day
53
in
the
50
mg/
kg/
day
group.
The
rats
were
mated
during
study
days
30­
37
and
49­
56
with
one
female
per
male,
and
during
study
days
65­
71
with
two
females
per
male.
No
effects
on
the
number
of
fertile
males,
litter
size,
fetal
body
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
49
Draft,
do
not
cite
or
quote
weight,
or
number
of
implants
per
litter
were
observed.
There
was
a
dose­
dependent,
but
not
always
statistically
significant,
reduction
in
copulating
pairs
and
copulatory
plugs
relative
to
controls
during
the
65­
71
day
mating
period.
During
this
final
mating
period,
there
was
also
a
dose­
dependent
decrease
in
the
number
of
males
siring
two
litters,
which
was
statistically
significant
only
at
the
highest
dose
tested,
50
mg/
kg/
day.
The
effects
on
mating
behavior
were
similar
at
10
mg/
kg/
day
and
50
mg/
kg/
day,
but
there
was
no
clear
dose
response.
For
example,

there
were
fewer
copulatory
plugs
and
multiple
litters
at
50
mg/
kg/
day
than
at
10
mg/
kg/
day,
but
the
higher
dose
had
more
copulatory
pairs
and/
or
inseminations
(
depending
on
the
mating
period).

The
mating­
behavior
effects
in
the
10
mg/
kg/
day
dose
group
included
fewer
copulating
pairs,

fewer
inseminations,
fewer
copulatory
plugs,
and
fewer
multiple
litters,
but
the
only
statistically
significant
(
p<
0.05)
effect
at
this
dose
was
fewer
copulating
pairs
in
the
Day
65­
71
group.

Artificial
insemination
of
LHRH­
synchronized
females
was
performed
with
sperm
from
an
ancillary
group
exposed
to
0,
2,
10,
or
50
mg/
kg/
day
and
sacrificed
on
Day
31
(
6
males/
group)
or
sacrificed
on
Day
79
(
10
males/
group).
No
significant
effects
on
reproductive
outcomes
were
observed.
Necropsy
results
revealed
that
caput
and
cauda
sperm
counts
were
significantly
reduced
(
p<
0.05)
at
the
high
dose
to
87%
and
70%
of
control
values,
respectively,
on
Day
31,
and
to
89%

and
82%
of
control
values,
respectively,
on
Day
79.
The
percent
motile
sperm
was
affected
on
Day
31,
but
not
on
Day
79
at
50
mg/
kg/
day.
Exposure
to
50
mg/
kg
resulted
in
moderate
changes
in
sperm
morphology,
including
head
and
tail
defects
and
fused
sperm,
when
examined
at
either
Day
31
or
Day
79.
No
gross
effects
on
sperm
morphology
were
seen
at
2
or
10
mg/
kg/
day.
A
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
50
Draft,
do
not
cite
or
quote
slight
increase
in
dead
fetuses
of
5%
(
compared
to
0%
in
the
controls)
was
reported
for
the
Day
79
artificial
insemination
group
dosed
with
50
mg/
kg
DBA.
No
other
statistically
significant,

developmentally­
toxic
effects
were
reported.

Histopathologic
results
from
the
Linder
et
al.
(
1995)
study
were
reported
separately
(
Linder
et
al.,
1997a).
Necropsies
of
rats
dosed
with
0
or
250
mg/
kg/
day
DBA
were
performed
24
hours
after
the
last
of
2,
5,
9,
16,
or
31
daily
doses.
A
necropsy
was
also
done
at
Day
228,

which
included
42
days
of
exposure
and
a
6­
month
recovery
period.
At
Days
2
and
5,
there
was
moderate
to
extensive
retention
of
Step
19
spermatids
(
normally
released
in
Stage
VIII)
in
Stage
IX
of
the
cycle
of
the
seminiferous
epithelium.
This
retention
was
also
seen
in
Stages
IX
and
X
at
Days
9
and
16,
and
in
virtually
all
Stage
IX
and
X
tubules
at
Day
31.
Also
at
Day
31,
there
was
retention
of
Step
19
spermatids
and
degenerating
Step
19
spermatids
in
Stage
XI
to
XIV
tubules.

Basally­
located
remnants
of
Step
19
nuclei
were
seen
in
Stages
X
to
XII
at
Day
5,
Stages
X
to
XII
at
Day
9,
Stages
IX­
XII
at
Day
16,
and
in
Stages
XI­
XIV
at
Day
31.
Fused
Step
19
spermatid
flagella
in
Stage
IX
were
seen
from
Day
5
through
Day
31.
Atypical
residual
bodies
were
seen
from
Day
5
through
Day
16
at
numerous
Stages,
and
appeared
in
the
epididymis
on
Days
9,
16,
and
31.
Debris
from
these
atypical
residual
bodies,
along
with
other
cytoplasmic
debris
from
the
testes,
was
observed
throughout
the
epididymis
on
Day
31.
Acrosomes
and/
or
heads
of
Step
12
and
later
spermatids
were
affected
at
Days
16
and
31,
and
maloriented
late
spermatids
were
also
observed
at
these
times.
Vacuolization
of
the
Sertoli­
cell
cytoplasm
was
seen
at
Day
31.
After
a
fairly
long
(
6
month)
recovery
period,
male
rats
dosed
with
250
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
51
Draft,
do
not
cite
or
quote
mg/
kg/
day
still
displayed
atrophic
seminiferous
tubules,
disorganization
of
sperm­
producing
tubules,
and
degeneration
of
mature
and
round
spermatids,
all
likely
resulting
from
effects
on
the
structure
and/
or
function
of
Sertoli
cells
and
indicative
of
permanent
damage
to
the
reproductive
system.

In
the
same
study,
necropsies
of
rats
given
0,
2,
10,
or
50
mg/
kg/
day
were
performed
24
hours
after
the
last
of
31
or
79
daily
doses.
Retention
of
Step
19
spermatids
near
the
tubule
lumen
in
Stage
IX
was
observed
at
Day
31
at
doses
of
10
mg/
kg/
day
and
higher.
At
Day
79
in
the
10
mg/
kg/
day
dose
group,
there
was
also
retention
of
Step
19
spermatids
in
Stages
IX­
XI
at
both
the
lumenal
and
basal
surfaces.
One
animal
at
this
dose
also
had
disorganized
and
atrophied
tubules,

but
this
effect
was
not
considered
to
be
treatment­
related
because
tubular
disorganization
was
not
seen
at
50
mg/
kg/
day.
The
50
mg/
kg/
day
dose
group
(
after
31
doses)
had
increased
retention
of
Step
19
spermatids,
with
moderate
numbers
observed
at
the
lumenal
surface
at
Stage
IX
and
at
the
basement
membrane
at
Stages
X­
XII.
Similar
effects
were
seen
when
this
dose
group
was
treated
for
79
days,
with
the
retention
of
Step
19
spermatids
also
occurring
at
Stages
IX
and
X.

Atypical
residual
bodies
were
present
in
Stage
IX
at
Day
31
and
occasionally
at
Stage
IX
and
other
stages
at
Day
79.
Cytoplasmic
debris
was
observed
throughout
the
epididymis
at
Day
79
in
the
50
mg/
kg/
day
dose
group.
No
histopathological
changes
were
detected
at
2
mg/
kg/
day
on
either
Day
31
or
Day
79.

Thus,
the
two
papers
for
this
study
(
Linder
et
al.,
1995;
Linder
et
al.,
1997a)
describe
increasingly
severe
effects
with
increasing
dose
and
increasing
exposure
duration.
No
adverse
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
52
Draft,
do
not
cite
or
quote
effects
were
seen
at
the
lowest
dose
tested,
2
mg/
kg/
day.
Dosing
with
250
mg/
kg/
day
for
as
few
as
8­
14
days
caused
decreased
mating
and
decreased
fertility,
and
adverse
reproductive
effects
were
only
partially
reversible
after
42
days
of
exposure
and
a
6­
month
recovery
period.
Overall,

this
study
identified
an
equivocal
LOAEL
of
10
mg/
kg/
day
and
a
corresponding
NOAEL
of
2
mg/
kg/
day
for
male
reproductive
effects,
based
on
histological
evidence
for
changes
in
seminiferous
tubule
staging
of
altered
spermatid
development.

Collectively,
the
studies
by
Linder
and
colleagues
have
used
a
number
of
different
experimental
protocols
to
investigate
the
effects
of
DBA
on
spermatogenesis
and
the
resulting
effects
on
male
fertility
(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b;
Linder
et
al.,
1995;
Linder
et
al.

1997a).
Based
on
the
results
of
these
studies,
DBA
was
clearly
spermatotoxic
in
rats
following
high­
dose
single
exposures
or
repeated
exposures
for
longer
periods
of
time
(
up
to
79
days).

Effects
on
spermatogenesis
were
the
most
sensitive
endpoint
because
they
were
observed
in
the
absence
of
other
toxicity
indicators.
Significant
changes
in
sperm
count,
morphology,
and
motility
were
generally
observed
at
doses
higher
than
those
associated
with
early
histopathologic
changes
in
spermatogenesis.
However,
the
susceptibility
of
humans
to
DBA­
induced
reproductive
toxicity
is
not
known
and
it
is
possible
that
rats
are
more
sensitive
than
humans.
In
the
absence
of
valid
and
reliable
human
data
on
the
relationship
between
sperm
quality
and
human
fertility,
and
on
the
relative
sensitivity
of
humans
versus
rats
to
DBA­
associated
reproductive
effects,
adverse
histopathologic
or
sperm­
quality
changes
in
rodents
are
considered
to
be
the
more
appropriate
choice
for
the
critical
effect
in
the
studies
by
Linder
and
his
colleagues
than
fertility
changes.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
53
Draft,
do
not
cite
or
quote
In
contrast
to
the
results
of
Linder
et
al.
(
1994a),
Vetter
et
al.
(
1998)
did
not
observe
a
significant
spermatotoxic
effects
of
acute
treatment
with
DBA.
Vetter
et
al.
(
1998)
evaluated
spermatotoxic
effects
of
DBA
as
a
positive
control
to
validate
a
computer­
assisted
semen
analysis
and
a
flow
cytometric
assay
for
cell­
membrane
integrity
as
alternatives
to
sperm­
motility
assays
for
assessing
male­
reproductive
toxicity.
Sexually­
mature
male
Crl:
CD(
SD)
BR
rats
(
4­
5/
group)

were
given
single
oral
doses
of
0,
600,
or
1200
mg/
kg
DBA
in
10
mL/
kg
deionized
water.
The
high
dose,
but
not
the
low
dose,
resulted
in
overt
toxicity.
The
rats
were
sacrificed
after
13
days,

at
which
time
vas­
deferens
sections
were
harvested
for
sperm
analysis,
and
sections
of
the
testes
and
epididymides
were
taken
for
histopathologic
analysis.
The
average
percent
motile
sperm
was
74.4%,
74.8%,
and
65.7%
for
the
control,
low­,
and
high­
dose
groups,
respectively.
The
average
percent
of
viable
sperm
was
90.9%,
91.4%,
and
88.7%,
for
the
control,
low­
and
high­
dose
groups,
respectively.
Neither
sperm
motility
nor
membrane
permeability
following
DBA
treatment
were
statistically
different
from
controls.
The
absence
of
an
effect
was
not
likely
due
to
general
failure
of
the
assays
since
a
second
positive
control,
 ­
chlorohydrin,
did
decrease
sperm
motility.

Following
DBA
treatment,
no
morphologic
changes
were
observed
in
the
sperm;
however,
large
basophilic
bodies
were
observed
in
the
testes
of
low­
dose
rats
(
3/
5
males)
and
high­
dose
rats
(
4/
4
males),
and
in
the
epididymides
of
high­
dose
rats
(
4/
4
males).
The
study
authors
noted
that
differences
in
experiments,
such
as
the
source
of
the
sperm
from
the
vas
deferens
versus
the
cauda
epididymis,
the
strain
of
rat,
and
the
time
of
sacrifice
(
13
versus
14
days
post­
treatment)
are
unlikely
explanations
for
the
absence
of
spermatotoxicity
in
this
study
as
compared
with
the
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
54
Draft,
do
not
cite
or
quote
results
reported
by
Linder
et
al.
(
1994a).
In
the
absence
of
further
data,
the
reasons
for
the
differing
results
in
the
Linder
et
al.
(
1994a)
and
Vetter
et
al.
(
1998)
experiments
remain
unresolved.
The
Vetter
et
al.
(
1998)
study
identified
a
LOAEL
of
600
mg/
kg/
day,
based
on
histopathologic
changes
in
the
testes.
A
NOAEL
could
not
be
determined.

Although
the
effects
of
DBA
on
male
reproductive­
tract
toxicity
have
been
well
studied,

fewer
studies
have
evaluated
the
potential
reproductive
effects
of
DBA
in
females.
Cummings
and
Hedge
(
1998)
studied
the
effects
of
DBA
exposure
during
early
pregnancy
in
rats.
Female
Holtzman
rats
(
8/
dose
group)
were
administered
gavage
doses
of
0,
62.5,
125,
or
250
mg/
kg/
day
DBA
dissolved
in
water
on
GD
1­
8.
Administration
of
a
higher
dose,
500
mg/
kg/
day,
induced
moribund
behavior
and
lethality;
therefore,
dosing
at
this
level
was
discontinued
and
these
animals
were
not
further
evaluated
for
reproductive
endpoints.
Treated
animals
from
the
other
dose
groups
were
sacrificed
on
GD
9,
and
body
and
reproductive­
organ
weights,
serum
levels
of
progesterone,
17 ­
estradiol,
and
luteinizing
hormone,
the
number
of
implantation
sites,
the
number
of
resorptions,
the
number
of
corpora
lutea,
and
pre­
implantation
losses
were
assessed.

The
only
affected
response
was
a
170
%
increase
in
serum
17 ­
estradiol
at
250
mg/
kg/
day.
A
second
group
of
females
was
dosed
similarly
to
the
first
group,
sacrificed
on
GD
20,
and
evaluated
for
body
weight,
preimplantation
losses,
number
of
resorptions,
number
of
pups
per
litter,
pup
weights,
and
placental
weights.
No
differences
in
any
of
these
measures
were
observed
between
treated
and
control
animals.
The
authors
concluded
that
DBA
had
little
effect
on
female
reproduction
for
the
measures
assessed
in
the
study,
but
noted
that
effects
on
ovarian
function
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
55
Draft,
do
not
cite
or
quote
future
fertility
were
not
tested;
such
tests
would
be
warranted
by
the
observed
increase
in
serum
17 ­
estradiol.
Based
on
the
increase
in
serum
17 ­
estradiol
in
this
study,
the
LOAEL
is
the
highest
dose
tested,
250
mg/
kg/
day,
and
the
NOAEL
is
125
mg/
kg/
day.
An
acute
FEL
of
500
mg/
kg/
day
was
also
identified,
based
on
moribund
behavior
and
lethality
in
the
pregnant
dams.

This
study,
however,
is
limited
by
the
small
sample
size
of
each
of
the
groups.

Christian
et
al.
(
2001)
evaluated
the
reproductive
and
developmental
toxicity
of
DBA
in
Sprague­
Dawley
rats.
Male
and
female
rats
(
10/
sex/
group)
were
given
DBA
in
deionized
drinking
water
at
concentrations
of
0,
125,
250,
500
or
1000
ppm,
beginning
14
days
prior
to
cohabitation
and
continuing
through
gestation
and
lactation
(
63­
70
days
of
treatment).
The
average
daily
doses
(
based
on
measured
water
consumption
and
body
weights)
varied,
depending
on
the
phase
of
reproduction.
For
males
throughout
the
study
(
SD
1­
70),
mean
daily
doses
were
10.2,
20.4,
35.7,

and
66.1
mg/
kg/
day,
respectively.
For
females
on
SD
1­
15
(
pre­
mating),
mean
daily
doses
were
13.3,
26.2,
41.8
and
60.2
mg/
kg/
day,
respectively;
and
14.8,
30.3,
48.5
and
81.6,
respectively,
on
gestation
day
(
GD)
0­
21.
During
lactation
(
LD
1­
29),
the
estimated
doses
were
0,
43.5,
86.6,

150.7
and
211.7
for
the
0,
125,
250,
500,
and
1000
ppm
groups,
respectively;
however,
these
doses
included
consumption
of
water
by
the
pups
and
thus
overestimated
the
mean
daily
intake
for
lactating
females.
Among
the
pups,
two
male
and
two
female
weanlings
from
each
litter
were
selected
for
one
additional
week
of
observation
(
postweanling
days
1­
8,
commencing
on
LD
29);

daily
food
intake,
drinking
water
consumption
and
body
weights
were
recorded,
and
necropsy
was
conducted
at
sacrifice.
The
mean
daily
doses
for
the
weanling
pups
were
0,
31.8,
58.5,
122.9
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
56
Draft,
do
not
cite
or
quote
and
254.7
mg/
kg/
day
for
males,
and
0,
33.3,
61.5,
123.8,
and
241.2
for
females
in
the
0,
125,
250,

500
and
1000
ppm
groups,
respectively.

Apparent
taste
aversion
was
associated
with
an
exposure­
dependent
reduction
in
water
consumption,
which
was
paralleled
by
a
reduction
in
food
intake
at
all
concentrations.
Decreased
body
weight
gain
was
observed
in
parental
animals
and
postweanling
pups
at
the
two
highest
exposure
levels.
Estrous
cycling
was
unaffected
in
the
female
rats.
The
only
observed
adverse
reproductive
effect
was
a
possible
reduction
in
mating
performance
in
the
1000
ppm
group,
as
evidenced
by
a
slight
but
nonsignificant
increase
in
the
number
of
days
of
cohabitation
and
a
decrease
in
the
number
of
mated
pairs
(
6/
10
in
the
1000
ppm
group
versus
9­
10/
10
in
all
other
groups).
There
were
no
effects
on
pre­
and
postimplantation
losses,
live
litter
sizes,
and
gross
external
morphology
or
sex
ratios
in
the
pups.
Although
an
exposure­
related
decrease
in
pup
body
weights
was
noted,
these
findings
were
attributed
to
decreased
water
and
food
consumption
resulting
from
the
poor
palatability
of
DBA­
treated
drinking
water.
Based
on
a
lack
of
statistically
significant,
treatment­
related
findings,
the
parental
and
reproductive/
developmental
NOAEL
for
this
study
is
the
highest
dose
tested,
and
a
LOAEL
could
not
be
determined.
For
males,
the
paternal
NOAEL
is
66
mg/
kg/
day;
for
females,
the
corresponding
NOAEL
is
not
less
than
60
mg/
kg/
day,
and
is
likely
to
be
higher,
as
water
consumption
and
corresponding
mean
DBA
daily
doses
were
increased
during
gestation
(
to
82
mg/
kg/
day)
and
lactation
(
mean
daily
doses
could
not
be
determined
due
to
the
confounding
effects
of
water
consumption
by
the
pups).
Similarly,

the
NOAEL
for
developmental
effects
is
at
least
82
mg/
kg/
day
(
maternal
dose
during
gestation).
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
57
Draft,
do
not
cite
or
quote
The
Chemistry
Council
(
CCC,
2001;
Christian
et
al.,
2002)
recently
completed
a
twogeneration
drinking
water
study
of
DBA
in
rats,
conducted
according
to
Good
Laboratory
Practice
(
GLP)
standards
and
U.
S.
EPA
test
guidelines.
The
report
has
recently
been
published
and
has
also
been
independently
reviewed
and
accepted
by
an
EPA
scientific
advisory
group.

Because
this
study
addresses
a
key
data
gap,
a
fairly
detailed
summary
of
the
reported
findings
is
presented
here.
Male
and
female
Crl:
CD
Sprague­
Dawley
rats
(
30/
sex/
exposure
group)
were
administered
DBA
in
drinking
water
at
concentrations
of
0,
50,
250,
or
650
ppm
continuously
from
initiation
of
exposure
of
the
parental
(
P)
generation
male
and
female
rats
through
weaning
of
the
F2
offspring.
The
concentrations
were
chosen
based
on
a
range­
finding
study
that
found
that
650
ppm
was
the
highest
concentration
expected
to
allow
survival
of
the
F1
offspring.
For
the
P
generation,
DBA
exposure
was
initiated
at
43
days
of
age
and
continued
from
premating
until
study
day
(
SD)
92
for
males;
and
from
premating
through
gestation
and
a
29­
day
period
of
lactation
(
LD
1­
29)
(
for
approximately
120
days
of
exposure)
for
females.
Parental
generation
offspring
(
F1
males
and
females)
were
exposed
in
utero
during
gestation,
and
during
lactation
(
LD
1­
29);
selected
F1
males
and
females
(
30/
sex/
exposure
group)
were
further
exposed
during
a
postweaning
period
of
at
least
71
days,
which
continued
through
mating,
gestation,
and
lactation.

All
other
F1
pups
were
sacrificed
on
LD
29.
All
F1
adult
females
and
their
offspring
(
F2
generation)
were
sacrificed
on
LD
22.
All
females
in
the
P
and
F1
generations
were
evaluated
once
daily
for
estrous
cycling
(
from
21
days
before
cohabitation
through
GD
0).
All
females
were
also
assessed
for
duration
of
gestation,
fertility
index,
gestation
index,
number
and
sex
of
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
58
Draft,
do
not
cite
or
quote
offspring
per
litter,
number
of
implantation
sites,
litter
size
and
viability,
viability
index,
lactation
index,
percent
pup
survival
and
litter
sex
ratio,
general
condition
of
the
dam
and
litter
during
the
postpartum
period,
and
maternal
behavior
during
lactation.
Litters
were
examined
to
identify
external
abnormalities,
physical
signs
of
toxicity,
pup
weights,
and
litter
viability.
Necropsy
of
all
P
and
F1
adults
included
gross
evaluation
of
the
cranial,
thoracic,
abdominal,
and
pelvic
viscera.

Specialized
measurements
evaluating
sperm
parameters
(
concentration,
percent
motility,

morphology,
number
of
sperm,
and
testicular
spermatid
count).
F1
generation
pups
were
also
evaluated
for
age
at
sexual
maturation
(
as
determined
by
vaginal
patency
in
females,
preputial
separation
in
males,
and
anogenital
distance
in
both
sexes)
were
performed.
Individual
organ
weights
were
recorded
for
major
organs,
including
testes
and
ovaries,
as
well
as
uterus
with
oviducts
and
cervix,
epididymides,
prostate
gland,
and
seminal
vesicles
with
coagulating
glands.

All
gross
lesions
were
examined
histologically.
Histopathology
was
also
conducted
on
adrenal
and
pituitary
glands;
testis,
epididymides,
prostate,
seminal
vesicles,
coagulating
glands
in
males;

ovaries,
oviducts,
uterus,
cervix
and
vagina
in
females;
and
selected
additional
organs
based
on
observed
organ
weight
changes.
Additionally,
testicular
histopathology
examining
the
caput,

corpus,
and
cauda
of
the
epidiymis
was
conducted
in
males,
and
the
reproductive
organs
of
all
rats
suspected
of
reduced
fertility
were
subjected
to
histological
examination.

The
average
daily
doses
(
based
on
measured
water
consumption
and
body
weights)
are
presented
in
Table
V­
10.
Daily
doses
varied
both
between
exposure
groups
and
among
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
59
Draft,
do
not
cite
or
quote
reproductive
stages
(
premating,
gestation,
lactation).
Significant
increases
in
pup
mortality
in
F1
litters
were
considered
to
be
unrelated
to
DBA
exposure
because
the
incidences
were
within
the
historical
control
of
the
testing
facility.
Other
unscheduled
deaths
in
the
study
were
also
unrelated
to
exposure
to
DBA.
Clinical
signs
of
toxicity
were
observed
in
various
groups
exposed
to
250
and
650
ppm,
and
included
soft
or
liquid
feces,
dehydration,
and
ungroomed
coats.
Water
consumption
were
statistically
significantly
decreased
in
the
P
and
F1
generation
at
all
exposure
levels,
presumably
due
to
taste
aversion,
and
food
intake
was
significantly
reduced
at
the
highest
dose
group
in
the
P
generation
and
the
two
high
exposure
groups
in
the
F1
generation.
Body
weights
and
body
weight
gains
for
high­
dose
P
males
and
females
were
significantly
reduced
during
the
premating
period
and
were
significantly
decreased
for
high­
dose
P
females
during
gestation
and
lactation.
F1
male
and
female
pups
had
significantly
reduced
body
weights
at
all
exposure
levels
during
the
lactation
period,
sufficient
for
the
study
authors
to
delay
weaning
until
LD
29
to
ensure
pup
survival.

Table
V­
10.
Average
Consumed
Daily
Doses
(
mg/
kg/
day)
for
Male
and
Female
Sprague­
Dawley
Rats
in
the
Two­
Generation
Reproductive/
Developmental
Toxicity
Studya
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
60
Draft,
do
not
cite
or
quote
DBA
Exposure
Groups
0
ppm
50
ppm
250
ppm
650
ppm
P
Generation
 
Male
Rats
Premating
to
Termination
(
SD
1­
92)
0.0
4.4
22.4
52.4
P
Generation
 
Female
Rats
Premating
to
Cohabitation
(
SD
1­
70)
0.0
6.0
28.1
69.1
Gestation
(
GD
0­
21)
0.0
6.4
30.1
76.1
Lactation
(
LD
1­
15)
0.0
11.6
55.6
132.0
F1
Generation
 
Male
Rats
Premating
(
PD
1­
71)
0.0
5.7
29.7
74.6
Weaning
to
Termination
(
PD
1­
134)
0.0
4.5
22.0
54.7
F1
Generation
 
Female
Rats
Weaning
to
Cohabitation
(
PD
1­
71)
0.0
6.6
32.1
83.4
Gestation
(
GD
0­
21)
0.0
6.2
28.5
67.1
Lactation
(
LD
1­
15)
0.0
10.0
49.6
114.7
a
Chlorine
Chemistry
Council
(
2001),
unpublished
report;
Christian
et
al.,
(
2002)
bSD
=
study
day;
GD
=
gestation
day;
LD
=
lactation
day;
PD
=
postweaning
day
By
LD
29,
the
body
weights
of
pups
in
the
50
ppm
group
were
similar
to
control
pup
weights.
Throughout
the
postweaning/
premating
period,
F1
males
and
females
in
the
250
and
650
ppm
groups
weighed
significantly
less
than
controls,
and
the
females
continued
to
exhibit
significant
reductions
in
body
weight
(
compared
to
controls)
during
gestation
and
lactation.
The
body
weights
of
F2
pups
in
the
two
highest
dose
groups
were
also
reduced
by
the
end
of
lactation;
however,
these
reductions
were
not
reported
to
be
statistically
significant,
relative
to
control
group
values.
No
treatment­
related
effects
were
reported
in
either
generation
for
estrous
cycling,
number
of
days
in
cohabitation,
duration
of
gestation,
mating
indices,
fertility
indices,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
61
Draft,
do
not
cite
or
quote
number
and
sex
of
offspring
per
litter,
number
of
implantation
sites,
litter
size,
lactation
index,

percent
pup
survival,
pup
sex
ratio,
and
gross
malformations.
Total
litter
loss
observed
in
the
P
generation
for
two
dams
in
the
250
ppm
exposure
group
and
one
dam
in
the
650
ppm
group
was
not
considered
to
be
treatment­
related.

For
the
F1
generation
650
ppm
exposure
group,
preputial
separation
was
significantly
delayed
in
the
male
rats
(
50.5
days
versus
48.1
days
in
controls),
and
vaginal
patency
in
female
rats
was
also
significantly
retarded
(
36.3
days
versus
33.4
days
in
controls);
no
significant
difference
was
seen
when
the
data
were
analyzed
using
body
weight
as
a
covariant.
These
effects
were
considered
to
be
due
to
a
general
retardation
of
growth
associated
with
the
significant
reduction
in
body
weight
in
this
exposure
group
at
weaning.
In
F2
male
and
female
pups,

anogenital
distance
did
not
differ
from
controls
on
LD
1
but
was
significantly
reduced
in
male
pups
in
the
250
and
650
ppm
by
LD
22;
these
findings
were
also
considered
to
be
associated
with
a
general
retardation
of
growth
rather
than
being
treatment­
related.

An
increased
incidence
of
malformations
of
the
male
reproductive
tract,
including
small
testes
and
small
or
absent
epididymides,
was
observed
in
four
males
in
the
F1
group
exposed
to
650
ppm
and
was
considered
to
be
treatment­
related.
Histomorphologic
examination
of
these
organs
in
these
males
revealed
a
minimal
increase
in
abnormal
residual
bodies,
retained
Step
19
spermatids,
hypospermia,
atrophied
epididymis
and/
or
atrophied
testis.
Histopathologic
examination
of
reproductive
organs
of
P
and
F1
male
rats
in
the
250
and
650
ppm
groups
(
N
=

30/
group/
generation)
showed
a
consistent
and
significant
exposure­
related
increase
in
retained
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
62
Draft,
do
not
cite
or
quote
Step
19
spermatids
in
Stage
IX
and
X
tubules
and
in
increased
and
abnormal
residual
bodies
in
affected
seminiferous
tubules
(
Table
V­
11).
Diffuse
testicular
atrophy
and
phagocytized
Step
19
nuclei
in
the
basilar
area
of
affected
seminiferous
tubules
were
also
observed,
although
at
a
lower
incidence.
Other
testicular
abnormalities
in
250
and
650
ppm
male
rats
of
both
generations
included
increased
amounts
of
exfoliated
spermatogenic
cells/
residiual
bodies
in
epididymal
tubules,
atrophy,
and
hypospermia.
Percent
motile
sperm,
sperm
count,
sperm
density,
and
number
and
percent
of
morphologically
abnormal
sperm
for
exposed
groups
were
within
historical
control
values
for
the
test
laboratory
and
were
unaffected
by
treatment.
No
effects
were
observed
in
the
prostate
gland,
seminal
vesicles,
or
coagulating
glands
of
any
of
the
male
rats
of
either
generation.
All
gross
lesions
in
other
organs
in
P
and
F1
parental
rats
and
in
F1
ands
F2
pups
were
considered
to
be
unrelated
to
DBA
treatment.

Table
V­
11.
Incidences
of
Exposure­
Related
Histopathologic
Findings
in
the
Testes
of
Rats
Consuming
DBA
in
Drinking
Water
a
P
F1
DBA
Concentrations
(
ppm)
0
50
250
650
0
50
250
650
Number
of
Testes
Examined
30
30
30
30
30
30
30
30
Retention
of
Step
19
Spermatids
4
3
13
23
0
1
12
20
Abnormal/
Increased
Residual
Bodies
3
5
15
25
1
2
10
14
a
Chorine
Chemistry
Council
(
2001),
unpublished
report
Histologic
examination
of
the
ovaries
of
ten
P
and
F1
female
rats
in
the
control,
250
and
650
ppm
exposure
groups
did
not
reveal
any
functional
abnormalities;
corpora
lutea
and
growing
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
63
Draft,
do
not
cite
or
quote
and
antral
follicles
were
present
and
apparently
normal.
There
were
no
significant
differences
in
the
number
of
postlactational
ovarian
primordial
follicles
among
any
of
these
groups.

A
variety
of
decreases
in
organ
weights
were
observed
that
were
attributed
to
general
growth
retardation.
In
addition,
increases
in
absolute
and
relative
kidney
and
liver
weights
(
of
approximately
10%)
were
observed
in
the
P
and
F1
males
and
females.
There
was
no
doseresponse
in
increased
kidney
weight,
although
the
increase
in
absolute
and
relative
liver
weight
was
dose­
related.
There
was
no
supporting
histopathology
in
an
evaluation
of
the
liver
and
kidney
in
10
rats/
sex/
group,
and
the
study
authors
did
not
consider
the
organ
weight
changes
to
be
toxicologically
significant.
Histopathology
in
the
zona
glomerulosa
of
the
adrenal
cortex
in
female
rats
of
all
DBA
exposure
groups
of
both
generations
was
considered
to
be
a
physiologic
response
related
to
water
balance
and/
or
stress,
and
not
a
direct
exposure­
related
effect.
F1
generation
male
and
female
rats
had
significant
increased
spleen
weights
relative
to
terminal
body
weights.

An
increase
in
the
incidence
and
intensity
of
extramedullary
hematopoiesis
in
the
red
pulp
of
the
spleen
occurred
in
the
F1
generation
female
rats
in
the
650
ppm
group
and
may
have
been
treatment­
related.
Decreased
cellularity
of
the
cortical
lymphoid
area
of
the
thymus
was
noted
in
P
generation
females
in
the
two
highest
dose
groups.

The
parental
NOAEL
for
general
toxicity
is
50
ppm,
based
on
increase
in
absolute
and
relative
liver
and
kidney
weights.
Based
on
testicular
histomorphology
indicative
of
abnormal
spermatogenesis
in
P
and
F1
males,
the
reproductive/
developmental
toxicity
LOAEL
and
NOAEL
are
250
and
50
ppm,
respectively.
For
the
P
generation,
these
drinking
water
concentrations
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
64
Draft,
do
not
cite
or
quote
correspond
to
a
LOAEL
and
NOAEL
of
22
and
4
mg/
kg/
day,
respectively.
For
the
F1
generation,

mean
daily
doses
are
considered
to
be
equivalent
to
the
mean
of
average
consumed
doses
during
the
period
from
weaning
to
termination
of
the
study.
These
doses
are
very
similar
to
those
for
the
P
generation;
resulting
in
a
LOAEL
and
NOAEL
for
the
F1
generation
of
22.0
and
4.5
mg/
kg/
day,
respectively,
equivalent
to
drinking
water
concentrations
of
250
and
50
ppm
in
drinking
water,
respectively.

The
developmental
toxicity
of
DBA
has
been
reported
in
two
related
abstracts.
Narotsky
et
al.
(
1996)
studied
the
developmental
toxicity
of
DBA
in
CD­
1
mice
dosed
by
gavage
with
0,

0.11,
0.23,
0.46,
0.92,
1.8,
2.8,
or
3.7
mmol/
kg/
day
(
equivalent
to
0,
24,
50,
100,
200,
392,
610,

and
806
mg/
kg/
day)
on
GD
6­
15.
Mice
were
allowed
to
deliver
naturally
and
the
litters
were
examined
on
postnatal
days
(
PND)
1
and
6.
Maternal
effects
were
limited
to
piloerection
and
motor
depression
at
the
highest
dose
tested.
Parturition
was
delayed
at
all
doses
tested
but
the
toxicologic
significance
of
this
effect
is
unclear.
In
the
highest­
dose
group
(
806
mg/
kg/
day),

prenatal
mortality
was
increased,
and
only
3/
9
litters
were
viable
at
birth.
Increased
postnatal
mortality
was
seen
at
610
and
806
mg/
kg/
day.
Decreased
pup
weight
was
observed
at
806
mg/
kg/
day
on
PND
1
and
at
610
mg/
kg/
day
on
PND
6.
Skeletal
malformations,
as
indicated
by
short,
kinked,
or
absent
tails,
were
in
the
two
highest­
dose
groups.
Based
on
these
results,
the
authors
concluded
that
DBA
was
a
developmental
toxicant.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
65
Draft,
do
not
cite
or
quote
In
a
second
published
abstract,
DBA
was
administered
to
CD­
1
mice
by
gavage
in
distilled
water
on
GD
6­
15
at
doses
of
0,
50,
100,
or
400
mg/
kg/
day
(
Narotsky
et
al.,
1997).
Maternal
toxicity
was
not
observed.
Litters
were
removed
by
cesarean
section
on
GD
17,
and
half
of
the
fetuses
in
each
litter
were
examined
for
skeletal
defects
and
the
other
half
for
soft­
tissue
malformations.
There
were
no
effects
on
prenatal
survival,
fetal
weight,
and
skeletal
development.

Hydronephrosis
was
noted
at
100
and
400
mg/
kg/
day,
and
renal
agenesis
(
small
kidneys)
was
observed
at
400
mg/
kg/
day.
In
contrast
to
the
Narotsky
et
al.
(
1996)
abstract,
which
reported
delayed
parturition
at
24
mg/
kg/
day
and
above,
the
second
abstract
showed
no
developmentallyadverse
effects
at
the
50
mg/
kg/
day
dose.
Based
on
the
summary
data
provided
in
these
two
abstracts,
the
LOAEL
for
fetal­
kidney
malformations
would
be
100
mg/
kg/
day,
with
a
corresponding
NOAEL
of
50
mg/
kg/
day.
However,
due
to
the
limited
data
provided
in
the
abstracts,
the
use
of
the
reported
adverse­
effect
levels
for
quantitative
risk
assessment
is
not
appropriate.

Klinefelter
et
al.
(
2000),
in
an
abstract,
reported
the
effects
of
DBA
administered
in
drinking
water
on
the
pubertal
development
and
adult
reproductive
function
of
male
Sprague­

Dawley
rats
(
3
litters/
dose)
exposed
from
GD
15
to
PND
98.
Pregnant
and
lactating
dams
were
exposed
to
0,
400,
600,
or
800
ppm
DBA
in
drinking
water,
equivalent
to
0,
50,
75,
and
100
mg/
kg/
day
(
personal
communication
with
authors).
After
weaning,
male
offspring
were
exposed
to
the
same
concentrations
of
DBA
in
drinking
water
and
sacrificed
on
PND
98.
Histologic
examination
of
the
reproductive
tract
was
performed
on
one­
half
of
the
sacrificed
animals;
the
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
66
Draft,
do
not
cite
or
quote
other
half
was
used
for
harvesting
of
proximal
cauda­
epididymis
sperm
for
artificial
insemination
of
LHRH­
synchronized
females.
Decreased
body
weight
throughout
the
reproductivedevelopment
period
was
observed
in
the
high­
dose
male
offspring
as
compared
with
control
animals.
Decreased
epididymis
weight
(
the
percent
decrease
was
not
specified)
occurred
in
the
75
and
100
mg/
kg/
day
groups.
The
age
at
preputial
separation
was
delayed
in
all
treatment
groups,

averaging
49,
48,
and
50
days
for
the
50,
75,
and
100
mg/
kg/
day
dose
groups,
respectively,
as
compared
with
42
days
in
controls.
Histopathologic
examination
revealed
the
presence
of
only
Sertoli
cells
in
the
seminiferous
tubules
of
animals
in
all
dose
groups.
The
fertility
of
treated
males
was
also
affected
by
DBA
treatment.
The
number
of
implants
per
corpora
lutea
in
females
artificially
inseminated
with
sperm
from
treated
males
decreased
from
70%
for
controls
to
49%,

15%,
and
15%
for
the
50,
75,
and
100
mg/
kg/
day
dose
groups,
respectively.
Levels
of
the
sperm
protein
SP22,
which
has
been
shown
to
be
highly
correlated
with
rodent
fertility,
were
significantly
decreased
in
all
treatment
groups.
Based
on
adverse
effects
on
the
fertility
of
sperm
of
treated
males,
the
lowest
dose
tested,
50
mg/
kg/
day,
would
be
the
LOAEL,
and
a
NOAEL
could
not
be
determined.
However,
due
to
the
limited
data
provided
in
this
abstract,
the
use
of
the
reported
adverse­
effect
levels
for
quantitative
risk
assessment
is
not
appropriate.
Further,

according
to
the
study
authors,
a
more
comprehensive
study
using
lower
doses
is
being
conducted
to
identify
the
NOAEL/
LOAEL
boundary
(
personal
communication).

In
a
second
recent
abstract,
Veeramachaneni
et
al.
(
2000)
exposed
male
Dutch­
belted
rabbits
(
10/
group)
to
DBA­
treated
drinking
water
from
GD15
throughout
life.
The
average
daily
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
67
Draft,
do
not
cite
or
quote
doses
were
reported
as
0,
0.97,
5.05,
and
54.2
mg/
kg/
day.
The
ability
of
the
treated
males
to
ejaculate
was
determined
by
collecting
ejaculates
every
3­
4
days,
beginning
at
20
weeks
of
age.

One
male
in
each
of
the
0.97
and
54.2
mg/
kg/
day
dose
groups
consistently
failed
to
ejaculate,
and
one
male
in
each
of
the
5.05
and
54.2
mg/
kg/
day
dose
groups
failed
to
ejaculate
at
least
once.
In
the
54.2
mg/
kg/
day
dose
group,
males
that
did
ejaculate
took
more
attempts
and
longer
time
to
ejaculate
compared
to
controls
(
p<
0.05).
The
fertility
of
sperm
from
24­
week­
old
males
was
assessed
by
artificial
insemination
of
two
6­
month­
old
rabbit
females
per
sample
of
sperm
from
each
male.
Conception
rates
were
significantly
decreased
(
p<
0.01)
in
females
inseminated
with
sperm
from
males
in
all
treated
groups,
averaging
85%,
55%,
65%,
and
55%
for
rabbit
does
inseminated
with
sperm
from
males
treated
with
0,
0.97,
5.05,
and
54.2
mg/
kg/
day
DBA,

respectively.
Of
the
53
pups
born
to
females
inseminated
with
sperm
from
the
high­
dose
males,
1
pup
had
cleft
palate
and
cranioschisis,
and
2
pups
had
cranioschisis.
At
25
weeks,
the
offspring
were
necropsied;
no
differences
in
body
weight,
anogenital
distance,
or
sex­
organ
weights
were
reported
relative
to
controls.
These
abstract
data
suggest
that
the
lowest
dose
tested,
0.97
mg/
kg/
day
was
a
LOAEL
for
decreased
male
fertility
and
that
a
NOAEL
could
not
be
determined.

However,
a
critical
assessment
of
these
findings
cannot
be
conducted
without
a
full
review
of
the
study
report.

Taken
together,
the
data
provide
strong
evidence
that
DBA
is
a
male
reproductive
system
toxicant
following
oral
dosing.
The
gavage
studies
of
Linder
and
colleagues
reported
perturbation
of
spermatogenesis
based
on
histopathology
changes
in
seminiferous­
tubule
staging,
changes
in
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
68
Draft,
do
not
cite
or
quote
sperm
quality
(
count,
morphology,
and
motility),
and
in
male
reproductive
performance
(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b;
Linder
et
al.,
1995;
Linder
et
al.
1997a).
Administration
of
DBA
in
drinking
water
has
also
been
reported
to
adversely
affect
both
sperm
quality
and
male
reproductive
performance
in
young
males
exposed
continuously
during
gestation,
lactation,
and
the
post­
weaning
developmental
period
(
Klinefelter
et
al.,
2000,
Veeramachaneni
et
al.,
2000).
In
contrast,
although
the
two­
generation
drinking
water
reproductive
toxicity
study
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002)
reported
testicular
histomorphology
indicative
of
abnormal
spermatogenesis
similar
to
that
found
in
shorter­
term
studies
by
Linder
and
colleagues,

no
adverse
treatment­
related
effects
on
mating
performance,
fertility,
gestation
length,
and
other
functional
indices
of
successful
reproductive
behavior
were
noted
at
mean
paternal
(
P
generation)

daily
doses
up
to
52
mg/
kg/
day,
and
at
mean
F1
daily
doses
up
to
55
mg/
kg/
day.

No
effect
on
female
reproductive
success
was
reported
in
Holtzman
rats
administered
DBA
doses
up
to
250
mg/
kg/
day
by
gavage
through
days
1
­
8
of
pregnancy
(
Cummings
and
Hedge,
1998).
Female
CD­
1
mice
given
gavage
doses
up
to
801
mg/
kg/
day
on
GD
6­
15
had
decreases
in
viable
litters,
increased
postnatal
mortality,
decreased
pup
weight,
and
increased
tail
abnormalities
(
Narotsky
et
al.,
1996).
In
a
second
study
examining
the
incidence
of
skeletal
and
visceral
malformations
in
the
pups
of
pregnant
CD­
1
mice
administered
DBA
gavage
doses
of
up
to
400
mg/
kg/
day,
an
increased
incidence
in
renal
malformations
(
Narotsky
et
al.,
1997)
was
reported
beginning
at
100
mg/
kg/
day.
Delayed
parturition
was
noted
at
all
doses
in
the
first,
but
not
the
second,
study
(
Narotsky
et
al.,
1996,
1997);
however,
details
were
not
reported
in
the
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
69
Draft,
do
not
cite
or
quote
abstracts
and
the
adversity
of
this
endpoint
is
unclear.
In
contrast,
no
treatment­
related
effects
on
litter
viability,
postnatal
mortality,
and
gross
malformations
were
observed
in
the
two­
generation
drinking
water
reproductive/
developmental
toxicity
study
(
Chlorine
Chemistry
Council,
2001;

Christian
et
al.,
2002);
as
previously
noted,
the
significant
decreases
in
the
body
weight
gain
in
pups
of
both
the
P
and
F1
generation
were
attributed
to
a
general
retardation
in
growth
associated
with
decreased
water
consumption
(
due
to
taste
aversion)
and
reduced
food
consumption,
not
to
a
direct
effect
of
DBA
treatment.
Differences
in
findings
between
the
two­
generation
study
and
those
by
Narotsky
et
al.
(
1996,
1997)
may
have
been
due
to
differences
in
internal
doses
associated
with
gavage
versus
drinking
water
DBA
administration,
species
differences
in
susceptibility
to
DBA
toxicity,
and/
or
the
lower
mean
doses
tested
in
the
two­
generation
study.

Gardner
and
Toussant
(
1999)
evaluated
developmental
toxicity
of
DBA
in
the
frog
embryo
teratogenesis
assay
­
Xenopus
(
FETAX)
(
a
96­
hour
toxicity
test),
with
and
without
metabolic
activation.
Endpoints
evaluated
were
embryolethality
(
LC
50),
embyronic
malformations
(
EC
50),
minimum
concentration
to
inhibit
growth
(
MCIG),
and
a
teratogenicity
index
(
TI
 
the
ratio
of
the
LC
50
to
the
EC
50).
The
FETAX
assay
is
considered
to
be
a
reliable
developmental
toxicity
screening
assay;
Dawson
and
Bantle
(
1987)
have
estimated
that
its
predictive
accuracy
for
identifying
known
mammalian
or
human
developmental
toxicants
approaches
or
exceeds
85%.

At
DBA
concentrations
of
up
to
12,800
mg/
L,
neither
50%
mortality
nor
50%
malformations
was
achieved
in
two
of
the
three
tests
conducted
without
added
metabolic
activation;
therefore,

neither
the
LC
50
nor
the
EC
50
could
be
estimated.
In
the
third
test,
the
LC
50
and
EC
50
without
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
70
Draft,
do
not
cite
or
quote
metabolic
activation
were
7,354
and
11,723
mg/
L,
respectively.
With
metabolic
activation,
the
LC
50'
s
for
three
tests
were
6,244,
69,
and
3,787
mg/
L;
the
reasons
for
the
low
LC
50
in
the
second
test
were
unclear
and
a
pooled
estimate
was
not
calculated.
The
EC
50,
estimated
for
one
test
only,

was
879
mg/
L.
The
TI
was
also
calculated
for
only
one
test,
and
was
0.6
with
metabolic
activation
and
0.1
without
metabolic
activation.
TI
values
of
>
1.5
suggest
teratogenic
potential;

therefore,
under
the
conditions
of
this
study,
DBA
did
not
exhibit
teratogenic
potential.
Further,

malformations
did
not
appear
to
increase
in
severity
or
prevalence
with
increasing
DBA
concentrations,
with
or
without
metabolic
activation.

No
data
were
identified
for
the
reproductive
or
developmental
toxicity
of
DBA
following
exposure
by
the
inhalation
or
dermal
route.

D.
Mutagenicity
and
Genotoxicity
Monobromoacetic
acid
MBA
induced
a
positive
mutagenic
response
in
Salmonella
typhimurium
in
the
standard
assay
system
(
NTP,
2000a).
Detailed
results
including
the
tester
strains
evaluated
and
microsomal
dependence
of
the
mutagenic
response
were
not
available
from
the
posted
testing
results.
Giller
et
al.
(
1997)
evaluated
the
mutagenicity
of
a
series
of
halogenated
acetic
acids,
including
monochloro,
dichloro,
trichloro,
monobromo,
dibromo,
and
tribromoacetic
acids
in
Salmonella
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
71
Draft,
do
not
cite
or
quote
typhimurium
strain
TA100
in
the
Ames­
fluctuation
test.
This
assay
is
a
modification
of
the
Ames
test
in
which
bacteria
are
exposed
to
the
compound
under
study
in
a
liquid
suspension.
Rather
than
determining
the
number
of
mutant
colonies,
the
fluctuation
assay
identifies
the
presence
of
mutants
based
on
a
change
in
color
of
the
liquid
medium
in
wells
containing
prototrophic
mutants.

MBA
was
tested
at
concentrations
of
0.03
to
30
µ
g/
mL
without
S9
activation,
and
at
0.3
to
300
µ
g/
mL
in
the
presence
of
S9
activation.
No
mutagenic
effect
was
detected
in
the
absence
of
S9
activation.
The
study
authors
indicated
that
10
µ
g/
mL
was
the
minimal
cytotoxic
dose
in
the
absence
of
S9
activation.
In
the
presence
of
S9
activation,
mutagenic
activity
was
observed
at
concentrations
ranging
from
20
to
75
µ
g/
mL.
The
decrease
in
positive
mutagenic
responses
at
the
high
doses
(
with
S9
metabolic
activation)
was
consistent
with
the
onset
of
cytotoxicity
at
100
µ
g/
mL.

Similar
results
were
reported
in
a
published
abstract
by
Kohan
et
al.
(
1998),
who
tested
the
mutagenicity
of
the
same
series
of
halogenated
acetic
acids
as
Giller
et
al.
(
1997).
S.

typhimurium
tester
strains
TA98
and
TA100
were
incubated
with
MBA,
with
or
without
S9
activation,
in
a
microsuspension
assay.
MBA
(
0.1
µ
mole)
induced
a
positive
mutagenic
response
in
both
strains
+
S9
at
subtoxic
concentrations
(
personal
communication).
Other
than
DBA
(
as
described
below),
none
of
the
other
halogenated
acetic
acids
induced
a
positive
mutagenic
response
when
tested
up
to
cytotoxic
concentrations.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
72
Draft,
do
not
cite
or
quote
Several
measures
of
DNA­
damage
response
have
been
reported
for
MBA.
Giller
et
al.

(
1997)
evaluated
DNA­
repair
responses
to
MBA
using
the
SOS
chromotest,
which
measures
the
induction
of
DNA
repair.
In
this
assay,
Escherichia
coli
strain
PQ37
was
exposed
to
concentrations
ranging
from
1
to
1000
µ
g/
mL
MBA
without
metabolic
activation,
and
from
3
to
3000
µ
g/
mL
with
metabolic
activation
by
S9
mix.
Toxic
concentrations
were
300
µ
g/
mL
and
higher,
regardless
of
S9
activation.
MBA
failed
to
induce
the
DNA­
repair
response
at
any
of
the
concentrations
tested,
regardless
of
metabolic
activation.

Giller
et
al.
(
1997)
also
evaluated
chromosome
damage
using
a
newt­
micronucleus
test.

Pleurodeles
waltl
larvae
were
exposed
to
varying
concentrations
of
MBA
in
the
absence
of
S9
for
12
days.
The
highest
concentration
tested
in
the
assay
was
half
the
minimum
concentration
that
led
to
detectable
physiological
disturbances
in
a
preliminary
test.
Fifteen
larvae
per
dose
group
were
exposed
to
10,
20,
or
40
µ
g/
mL
MBA
(
renewed
daily)
and
the
number
of
micronucleated
erythrocytes
in
a
sample
of
1000
erythrocytes
was
determined.
MBA
did
not
increase
the
number
of
micronuclei
at
any
of
the
tested
concentrations.

Stratton
et
al.
(
1981)
reported
that
MBA
concentrations
of
100
µ
M
(
13.9
mg/
L)
induced
DNA­
strand
breaks
in
L­
1210
mouse
leukemia
cells
as
measured
in
an
alkaline
elution
assay.

MBA
was
added
to
the
cell
culture
medium
and
the
cells
were
incubated
in
the
treated
medium
for
1
hour
in
the
absence
of
S9.
The
cells
were
rinsed
and
harvested
immediately
or
incubated
in
MBA­
free
medium
for
1
or
6
hours
before
measuring
DNA­
strand
breaks.
The
number
of
DNAstrand
breaks
was
increased
compared
to
controls
immediately
after
the
1­
hour
treatment,
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
73
Draft,
do
not
cite
or
quote
increased
even
further
following
post­
treatment
incubations.
The
study
authors
suggested
that
the
observed
increase
in
DNA­
strand
breaks
is
consistent
with
depurination
of
alkylated
DNA
over
time
to
form
alkali­
labile
apurinic
DNA
sites,
suggesting
that
MBA
can
induce
direct
DNA
damage.

The
results
of
genotoxicity
studies
for
MBA
are
summarized
in
Table
V­
12.
Based
on
these
limited
studies,
it
remains
unclear
if
MBA
is
genotoxic.
A
positive
mutagenic
response
was
observed
in
the
Ames
assays.
In
the
single
study
with
sufficient
detail
for
full
evaluation,
the
positive
finding
only
with
S9
activation
suggests
metabolic
activation
of
MBA
to
a
genotoxic
form,
but
there
have
been
no
metabolism
studies
to
identify
potentially
mutagenic
metabolites.

The
observed
effect
in
that
study
is
unlikely
to
be
due
to
altered
pH,
since
the
mutagenicity
was
observed
in
the
absence
of
cytotoxicity.
Although
MBA
was
mutagenic
in
the
Ames
assay,
it
did
not
induce
a
DNA
repair
response
in
the
SOS
chromotest.
In
addition,
measures
of
DNA
damage,

including
micronuclei
and
DNA­
strand
breaks,
have
yielded
inconsistent
results.
Taken
together,

these
data
are
not
sufficient
to
conclude
that
MBA
is
genotoxic.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
74
Draft,
do
not
cite
or
quote
Table
V­
12.
Genotoxicity
Studies
of
MBA
Endpoint
Assay
system
Results
(
wo/
w
activation)
Comments
Reference
Gene
mutationbacteria
Salmonella
typhimurium
+
Detailed
results
were
not
available
NTP,
2000a
Salmonella
typhimurium
TA100
­/+
Tested
to
cytotoxic
doses
in
Ames
fluctuation
protocol
Giller
et
al.,
1997
Salmonella
typhimurium
TA98,
TA100
­/+
Positive
in
TA98
and
TA100
in
suspension
assay.
Data
provided
in
a
published
abstract
Kohan
et
al.,
1998
Clastogenicity
Micronuclei
in
Pleurodeles
waltl
larvae
(
Newt)
erythrocytes
­/
NT
None
Giller
et
al.,
1997
DNA
damage
Escherichia
coli
strain
PQ37
SOS
chromotest
­/­
Tested
to
cytotoxic
doses
Giller
et
al.,
1997
Single
strand
breaks
in
L­
1210
mouse
leukemia
cells
+/
NT
MBA
was
shown
to
be
cytotoxic
to
L­
1210
cells
from
50

M.
Stratton
et
al.,
1981
NT
=
Not
tested
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
75
Draft,
do
not
cite
or
quote
Bromochloroacetic
acid
BCA
induced
a
positive
mutagenic
response
in
Salmonella
typhimurium
in
the
standard
assay
system
(
NTP,
2000b).
Detailed
results,
including
the
tester
strains
evaluated
and
microsomal
dependence
of
the
mutagenic
response,
were
not
available
from
the
posted
testing
results,
but
should
be
released
upon
completion
of
the
full
cancer
bioassay.
Two
related
studies
evaluated
the
ability
of
BCA
to
induce
oxidative
DNA
damage.
Austin
et
al.
(
1996)
investigated
the
hypothesis
that
compounds
that
induce
lipid
peroxidation
might
show
increased
potential
as
genotoxic
agents.
The
capacity
of
a
series
of
haloacetic
acids,
including
BCA
and
DBA
(
DBA
results
described
below),
to
induce
lipid
peroxidation
was
measured
by
an
increase
in
production
of
thiobarbituric
acid­
reactive
substances
(
TBARS)
in
the
liver.
As
an
indicator
of
genotoxicity,

oxidative
DNA
damage
in
the
liver
was
measured
by
an
increase
in
8­
hydroxydeoxyguanosine
(
8­

OHdG)
levels.
Male
B6C3F1
mice
(
number
per
group
varied
from
3
to
6)
were
exposed
to
single
oral
doses
of
0,
30,
100,
or
300
mg/
kg
BCA
by
gavage
in
distilled
water.
In
a
time­
course
experiment,
mice
were
given
300
mg/
kg
BCA
and
livers
were
harvested
at
1,
3,
5,
7,
9,
and
12
hours
after
dosing
for
measurement
of
liver
TBARS.
TBARS
levels
peaked
at
3
hours
after
dosing
and
reached
levels
approximately
5­
fold
greater
than
background.
TBARS
levels
returned
to
pre­
exposure
levels
between
7
and
9
hours
after
dosing.
BCA
at
300
mg/
kg
increased
8­
OHdG
levels
to
a
maximum
of
2­
to
3­
fold
above
controls
over
a
period
of
12
hours.
Both
TBARS
and
8­
OHdG
levels
increased
with
increasing
dose
when
measured
at
3
hours,
and
were
maximal
at
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
76
Draft,
do
not
cite
or
quote
the
highest
dose
tested.
For
both
TBARS
and
8­
OHdG,
the
increases
were
significant
(
p<
0.05)

beginning
at
30
mg/
kg.

Parrish
et
al.
(
1996)
evaluated
whether
the
ability
of
brominated
acetic
acids
to
induce
oxidative
stress
responses
was
due
to
peroxisome
proliferation.
The
effects
of
BCA
on
oxidative
DNA
damage
and
peroxisome
proliferation
were
measured
in
the
livers
of
male
B6C3F1
mice.

The
animals
(
6/
treatment
group)
were
given
drinking
water
containing
0,
100,
500,
or
2000
mg/
L
BCA
for
3
weeks.
The
approximate
corresponding
doses,
calculated
using
the
default
waterintake
value
of
0.25
L/
kg/
day
(
U.
S.
EPA,
1988),
are
25,
125,
and
500
mg/
kg/
day.
No
doserelated
change
in
body
weight
was
observed,
but
absolute
and
relative
liver
weight
increased
at
the
high
dose.
Two
responses
indicative
of
peroxisome
proliferation
(
increased
cyanide
insensitive
Acyl­
CoA
oxidase
activity
and
increased
12­
hydroxylation
of
lauric
acid)
were
also
studied,

because
peroxisome
proliferation
has
been
linked
with
the
hepatocarcinogenic
effect
of
trichloroacetate.
An
additional
dose
group
exposed
to
3000
mg/
L
BCA
(
750
mg/
kg/
day)
was
evaluated
for
the
Acyl­
CoA
activity
measurements.
BCA
had
no
effect
on
either
measure
of
peroxisome
proliferation
after
exposures
up
to
3000
mg/
L.
BCA
did
induce
oxidative
DNA
damage,
with
8­
OHdG
levels
in
nuclear
DNA
of
the
liver
significantly
increased
(
p<
0.05)

beginning
at
the
lowest
dose,
25
mg/
kg/
day.
The
level
of
8­
OHdG
increased
to
a
maximum
of
approximately
2­
fold
at
the
highest
dose
(
500
mg/
kg/
day).
It
is
not
clear
at
this
time
whether
the
parent
BCA
or
one
or
more
of
BCA
metabolites
is
responsible
for
the
observed
increase
in
oxidative
stress.
The
lack
of
correlation
of
8­
OHdG
levels
with
Acyl­
CoA
activity
or
12­
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
77
Draft,
do
not
cite
or
quote
hydroxylation
of
lauric
acid
suggests
that
peroxisome
proliferation
is
not
causally
associated
with
BCA­
induced
oxidative
stress.

The
results
of
Austin
et
al.
(
1996)
and
Parrish
et
al.
(
1996)
do
not
provide
evidence
of
a
direct
genotoxic
effect
of
BCA,
although
these
results
coupled
with
the
positive
results
in
the
Ames
assay
suggest
that
BCA­
induced
oxidative
stress
might
result
in
downstream
genotoxicity
through
oxidative
DNA
damage.

Dibromoacetic
acid
DBA
induced
a
positive
mutagenic
response
in
Salmonella
typhimurium
in
the
standard
assay
system
(
NTP,
2000c).
Detailed
results,
including
the
tester
strains
evaluated
and
microsomal
dependence
of
the
mutagenic
response,
were
not
available
from
the
posted
testing
results,
but
should
be
released
upon
completion
of
the
full
cancer
bioassay.
Giller
et
al.
(
1997)

evaluated
mutagenicity
of
a
series
of
halogenated
acetic
acids,
including
DBA,
in
the
Ames
fluctuation
test
as
described
previously
for
MBA.
DBA
was
tested
at
concentrations
ranging
from
3
to
3000
µ
g/
mL
without
S9
activation,
and
from
10
to10,000
µ
g/
mL
in
the
presence
of
S9
fraction.
Toxic
concentrations
were
1000
µ
g/
mL
without,
and
10,000
µ
g/
mL
with,
metabolic
activation.
Genotoxicity
of
DBA
was
detected
at
concentrations
of
10
to
750
µ
g/
mL
without
activation,
and
at
30
to
3000
µ
g/
mL
with
activation.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
78
Draft,
do
not
cite
or
quote
Similar
results
were
reported
for
the
microsuspension
Ames
assay
of
DBA
reported
in
a
published
abstract
by
Kohan
et
al.
(
1998).
DBA
(
2.0
µ
mole)
induced
a
positive
mutagenic
response
in
strains
TA98
and
TA100
+
S9
at
subtoxic
concentrations
(
personal
communication).

The
absence
of
a
positive
response
without
S9
activation
contrasts
with
the
report
of
Giller
et
al.

(
1997).

Saito
et
al.
(
1995)
analyzed
indoor
swimming­
pool
water
from
four
pools
for
the
presence
of
trace
halogenated
contaminants,
and
also
for
mutagenicity
using
the
Ames
Salmonella
typhimurium
assay
with
strains
TA98
and
TA100,
with
and
without
metabolic
activation.
As
part
of
the
study,
the
mutagenicity
of
DBA
(
reported
90%
purity)
was
also
investigated
in
strains
TA
98
and
TA100.
DBA
was
mutagenic
with
and
without
metabolic
activation
in
strain
TA100,
at
a
minimum
positive
concentration
of
640
µ
g/
plate
in
each
assay.
No
mutagenic
activity
was
identified
in
strain
TA98.

The
ability
of
DBA
to
induce
DNA­
repair
responses
has
been
evaluated
in
two
separate
reports.
Giller
et
al.
(
1997)
tested
the
ability
of
DBA
to
induce
DNA
damage
using
the
SOS
chromotest
as
described
previously
for
MBA.
E.
coli
strain
PQ37
was
exposed
to
10
to
10,000
µ
g/
mL
DBA
without
metabolic
activation,
and
to
3
to10,000
µ
g/
mL
with
S9
metabolic
activation.

Toxic
concentrations
were
reported
as
1000
µ
g/
mL
without
and
10,000
µ
g/
mL
with
S9
activation.
DBA
induced
a
positive
response
regardless
of
metabolic
activation.
The
concentrations
that
induced
DNA
repair
were
200
to
750
µ
g/
mL
without
activation,
and
100
to
3000
µ
g/
mL
with
activation.
Mayer
et
al.
(
1996)
presented
a
scheme
for
the
concentration
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
79
Draft,
do
not
cite
or
quote
analysis
of
water
samples
for
trace
analytes,
and
coupled
it
with
the
umu
Microtest,
which
measures
induction
of
DNA
repair.
DBA
was
positive
in
the
umu
Microtest,
with
and
without
metabolic
activation.
Details
of
this
report
were
in
German,
and
were
not
available
in
English
for
a
more
thorough
review.

As
described
above
for
BCA,
Austin
et
al.
(
1996)
tested
the
capacity
of
a
series
of
haloacetic
acids,
including
BCA
and
DBA,
to
induce
lipid
peroxidation
and
oxidative
DNA
damage.
Male
B6C3F1
mice
were
exposed
to
single
oral
doses
of
0,
30,
100,
or
300
mg/
kg
DBA
by
gavage
in
distilled
water.
In
a
time­
course
study,
mice
were
given
300
mg/
kg
DBA,
and
livers
were
harvested
at
1,
3,
5,
7,
9,
and
12
hours
after
dosing
for
measurement
of
liver
TBARS.

TBARS
levels
peaked
rapidly,
1
hour
after
dosing
for
DBA,
to
levels
approximately
5­
fold
greater
than
background,
and
returned
to
pre­
exposure
levels
between
7
and
9
hours
after
dosing.

DBA
at
300
mg/
kg
rapidly
increased
8­
OHdG
levels
2­
to
3­
fold.
In
contrast
to
TBARS,
the
increase
in
8­
OHdG
levels
was
sustained
over
a
12­
hour
period.
Both
TBARS
and
8­
OHdG
levels,
when
measured
at
1
hour,
increased
with
increasing
dose
and
were
maximal
at
300
mg/
kg,

the
highest
dose
tested.
For
TBARS,
the
increases
were
significant
(
p<
0.05)
beginning
at
300
mg/
kg,
and
increases
in
8­
OHdG
levels
were
significantly
greater
than
controls
beginning
at
30
mg/
kg.

Parrish
et
al.
(
1996)
tested
whether
the
ability
of
brominated
acetic
acids
to
induce
oxidative
stress
responses
was
due
to
peroxisome
proliferation.
The
effects
of
DBA
on
oxidative
DNA
damage
and
peroxisome
proliferation
were
measured
in
the
livers
of
male
B6C3F1
mice.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
80
Draft,
do
not
cite
or
quote
The
animals
(
6/
group)
were
given
drinking
water
containing
100,
500,
and
2000
mg/
L
DBA.
The
approximate
doses
calculated
from
a
default
water­
intake
value
of
0.25
L/
kg/
day
are
25,
125,
and
500
mg/
kg/
day
(
U.
S.
EPA,
1988).
No
dose­
related
change
in
body
weight
was
observed,
but
absolute
and
relative
liver
weight
increased
at
the
mid­
and
high
dose
for
DBA.
Two
responses
indicative
of
peroxisome
proliferation
(
increased
cyanide­
insensitive
Acyl­
CoA
oxidase
activity
and
increased
12­
hydroxylation
of
lauric
acid)
were
also
studied,
because
peroxisome
proliferation
has
been
linked
with
the
hepatocarcinogenic
effect
of
trichloroacetate.
An
additional
dose
group
exposed
to
3000
mg/
L
DBA
(
750
mg/
kg/
day)
was
evaluated
for
the
Acyl­
CoA
activity
measurements.
DBA
induced
Acyl­
CoA
activity
to
a
maximum
of
3­
fold
after
exposures
up
to
3000
mg/
L,
but
did
not
induce
the
12­
hydroxylation
of
lauric
acid.
The
study
authors
did
not
explain
the
inconsistency
in
the
different
responses
obtained
with
these
two
measures
of
peroxisome
proliferation.
DBA
induced
oxidative
DNA
damage,
with
8­
OHdG
levels
in
hepatic
nuclear
DNA
significantly
increased
(
p<
0.05)
at
the
highest
dose
(
500
mg/
kg/
day)
to
a
maximum
of
approximately
twice
the
control
response.
The
overall
lack
of
correlation
of
8­
OHdG
levels
with
Acyl­
CoA
activity
or
12­
hydroxylation
of
lauric
acid
suggests
that
peroxisome
proliferation
is
not
causally
associated
with
the
oxidative
stress
induced
by
brominated
acetic
acids.

Effects
of
DBA
have
also
been
evaluated
at
the
chromosome
level
in
one
study.
Giller
et
al.
(
1997)
conducted
the
newt­
micronucleus
test
for
DBA,
as
described
previously
for
MBA.

None
of
the
DBA
concentrations
that
were
tested
(
20,
40,
or
80
µ
g/
mL,
in
the
absence
of
S9)

significantly
increased
the
number
of
erythrocytes
with
micronuclei.
The
co­
clastogenic
effects
of
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
81
Draft,
do
not
cite
or
quote
various
water
pollutants
on
chromosomal
aberrations
induced
by
mitomycin
C
in
various
mammalian­
cell
lines
were
reported
by
Sasaki
and
Kinae
(
1995).
The
primary
focus
of
the
report
was
on
the
co­
clastogenic
effects
of
toxic
metals
such
as
lead
and
mercury;
however,
some
organic
chemicals
were
also
tested.
Post­
treatment
with
DBA
(
microsomal
activation
status
not
available
in
the
English
summary)
at
concentrations
up
to
approximately
15
µ
g/
mL
resulted
in
a
strong
dose­
related
increase
in
chromosomal
aberrations
induced
by
mitomycin
C.
Details
of
this
report
(
including
a
complete
description
of
the
test
system)
were
in
Japanese
and
were
not
available
in
English
for
further
review.

Of
the
brominated
acetic
acids,
the
database
for
DBA
is
most
complete,
as
summarized
in
Table
V­
13.
DBA
has
provided
nearly
uniformly­
positive
results
in
the
assays
tested.
The
positive
effects
have
been
reported
regardless
of
S9
activation,
suggesting
that
mutagenicity
is
independent
of
metabolism
by
cytochrome
P450s,
similar
to
DCA
whose
metabolism
does
not
involve
microsomal
activation
but
is
mediated
by
NADPH
and
GSH
(
Lipscomb
et
al.,
1995;

Cornett
et
al.,
1997;
Stacpoole,
1998).
DNA
damage
secondary
to
generation
of
oxidative
stress
has
been
reported
by
Austin
et
al.
(
1996),
and
is
likely
to
be
independent
of
peroxisome
proliferation
(
Parrish
et
al.,
1996).
The
induction
of
DNA­
damage
responses,
including
SOS
repair
system
(
Giller
et
al.,
1997)
and
the
umu
microtest
(
Mayer
et
al.,
1996),
supports
the
potential
mutagenicity
of
DBA.
On
the
other
hand,
no
induction
of
micronucleated
erythrocytes
was
reported
(
Giller
et
al.,
1997),
suggesting
that
DBA
was
not
clastogenic
in
the
newt
test
system.
No
evaluation
of
micronuclei
has
been
reported
in
the
more
standard
mouse
micronucleus
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
82
Draft,
do
not
cite
or
quote
assay.
The
clastogenicity
of
DBA
has
not
been
reported
in
other
assays
using
a
standard
protocol,

but
DBA
has
been
reported
to
be
co­
clastogenic
(
Sasaki
and
Kinae,
1995).
As
a
whole,
these
data
support
the
conclusion
that
DBA
is
mutagenic
and
genotoxic,
although
the
nature
of
the
DNA
damage
induced
by
DBA
remains
unclear.

Table
V­
14
provides
a
summary
of
the
genotoxicity
data
for
MBA,
BCA,
and
DBA.
The
data
are
inadequate
for
determining
whether
MBA
or
BCA
are
genotoxic,
but
suggest
that
DBA
is
genotoxic.
The
mechanism
by
which
these
different
brominated
acetic
acids
might
lead
to
DNA
damage
is
not
clear
from
these
data.
The
mutagenicity
of
MBA,
but
not
DBA,
might
be
metabolism
dependent.
The
data
are
very
sparse
for
BCA,
but
for
the
single
endpoint
evaluated,

BCA
and
DBA
shared
the
ability
to
induce
oxidative
DNA
damage.
Thus,
this
mechanism
remains
a
viable
explanation
for
the
onset
of
DNA
damage,
and
perhaps
mutagenicity
of
DBA
and
BCA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
83
Draft,
do
not
cite
or
quote
Table
V­
13.
Genotoxicity
Studies
of
DBA
Endpoint
Assay
system
Results
(
wo/
w
activation)
Comments
Reference
In
vitro
assays
Gene
mutationbacteria
Salmonella
typhimurium
+
Detailed
results
were
not
available
NTP,
2000c
Salmonella
typhimurium
TA100
+/+
Tested
to
cytotoxic
doses
in
Ames
fluctuation
protocol
Giller
et
al.,
1997
Salmonella
typhimurium
TA98,
TA100
­/+
Positive
in
TA98
and
TA100
in
suspension
assay.
Data
provided
in
a
published
abstract
Kohan
et
al.,
1998
Salmonella
typhimurium
TA98
and
TA100
+/+
Positive
in
TA
100.
Reference
in
Japanese;
only
study
summary
in
English
was
reviewed.
Saito
et
al.,
1995
Clastogenicity
Micronuclei
in
Pleurodeles
waltl
larvae
(
newt)
erythrocytes
­/
NT
None
Giller
et
al.,
1997
Not
specified.
+
Co­
clastogenic
with
mitomycin
C.
Reference
in
Japanese;
only
study
summary
in
English
was
reviewed.
Sasaki
and
Kinae,
1995
DNA
damage
Escherichia
coli
strain
PQ37
SOS
chromotest
+/+
Tested
to
cytotoxic
doses
Giller
et
al.,
1997
umu
Microtest
+/+
Reference
in
German;
only
study
summary
in
English
was
reviewed.
Mayer
et
al.,
1996
In
vivo
assays
DNA
damage
Oxidative
DNA
damage
mouse
liver
in
vivo
+
2­
to
3­
fold
induction
in
8OHdG
levels
Austin
et
al.,
1996
Oxidative
DNA
damage
mouse
liver
in
vivo
+
2­
to
3­
fold
induction
in
8OHdG
levels
Parrish
et
al.,
1996
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
84
Draft,
do
not
cite
or
quote
Table
V­
14.
Summary
of
Genotoxicity
Data
for
Brominated
Acetic
Acids
MBA
BCA
DBA
Assay
­
S9
+
S9
­
S9
+
S9
­
S9
+
S9
Mutagenicity
­
+
+
±
a
+

SOS
DNA
repair
induction
­
­
NTb
NT
+
+

DNA
damage
±
+
+

Chromosome
damage
­
NT
±
a.
Mixed
or
equivocal
results
are
denoted
with
a
±
.

b.
NT
=
not
tested.

E.
Carcinogenicity
Concern
for
the
potential
carcinogenic
hazard
of
the
brominated
acetic
acids
is
based
on
the
tumorigenicity
of
chlorinated
acetic
acids
observed
in
rodent­
cancer
bioassays
(
Boorman
et
al.,
1999).
Carcinogenicity
testing
data
for
the
brominated
acetic
acids
are
limited
to
results
reported
in
published
abstracts,
although
both
BCA
and
DBA
have
been
slated
for
complete
2­

year
cancer
bioassays
(
NTP,
2000b;
NTP,
2000c).

So
and
Bull
(
1995)
reported
in
a
published
abstract
that
DBA
increased
the
numbers
of
aberrant
crypt
foci
in
the
colon
of
F344
rats.
Male
rats
were
administered
an
initiating
dose
of
azoxymethane
and
were
exposed
to
1000
mg/
L
DBA
in
drinking
water
for
up
to
20
weeks.
The
number
of
aberrant
crypt
foci
and
the
complexity
of
the
foci
were
increased
in
the
animals
given
DBA
as
compared
to
animals
given
the
initiating
compound
only.
These
findings
may
be
of
particular
significance
because
the
colon
has
been
implicated
as
a
potential
cancer
site
in
humans
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
85
Draft,
do
not
cite
or
quote
exposed
to
disinfectant
by­
products
(
Boorman
et
al.,
1999).
Stauber
et
al.
(
1995),
in
an
abstract,

reported
that
preliminary
data
suggest
that
BCA
and
DBA
induce
hepatic
tumors
in
B6C3F1
mice.
No
experimental
details
were
provided
in
the
brief
summary.

F.
Summary
Monobromoacetic
acid
The
toxicity
data
for
MBA
are
very
limited.
The
oral
LD
50
for
MBA
was
reported
as
177
mg/
kg
in
male
rats
(
Linder
et
al.,
1994a).
MBA
is
a
dermal
irritant
when
topically
applied
to
the
skin
of
rabbits
(
Eriksson
et
al.,
1994).
The
systemic
toxicity
of
MBA
has
not
been
well
studied
by
any
route
of
exposure.
Reproductive­
toxicity
studies
are
limited
to
a
single­
dose
or
14­
day
oral
gavage
study
assessing
MBA
spermatotoxicity
(
Linder
et
al.,
1994a),
and
have
not
demonstrated
either
general
toxicity
or
spermatoxicity.
A
published
abstract
(
Randall
et
al.,
1991)
reported
decreased
maternal
weight
gain,
decreased
live­
fetus
size,
and
an
increased
incidence
of
softtissue
malformations
in
female
rats
orally
exposed
to
MBA
on
GD
6­
15;
the
LOAEL
and
NOAEL
for
these
effects
were
100
and
50
mg/
kg/
day,
respectively.
However,
the
full
study
has
not
been
published
and,
thus,
these
data
are
of
limited
utility
in
both
hazard
characterization
and
risk
assessment.
The
carcinogenicity
of
MBA
has
not
been
evaluated
by
any
route
of
exposure.
The
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
86
Draft,
do
not
cite
or
quote
genotoxicity
data
base
is
limited
to
four
in
vitro
and
one
in
vivo
newt
larvae
studies
the
results
of
which
are
mixed.

Bromochloroacetic
acid
The
database
for
BCA
toxicity
is
limited.
BCA
is
predicted
to
be
a
severe
dermal
irritant,

based
on
QSAR
modeling
(
Eriksson
et
al.,
1994).
Several
oral­
toxicity
studies
of
BCA
have
identified
the
kidney
and
liver
as
target
organs
of
systemic
toxicity,
although
reported
effects
have
been
minimal
and/
or
equivocal
(
NTP,
1998;
Austin
et
al.,
1996).
Although
BCA
did
not
induce
male
reproductive­
organ
toxicity
or
affect
sperm
quality
or
male
fertility
in
rats
in
an
NTP
(
1998)

reproductive
and
developmental
screening
assay,
three
more
recent
studies
reported
the
occurrence
of
reduced
sperm
quality
and
decreased
male
fertility
in
mice
and
rats.
In
a
published
abstract
by
Luft
et
al
(
2000),
male
mice
treated
with
72
mg/
kg/
day
BCA
for
14
days
had
impaired
sperm
quality
and
reduced
fertility;
no
effects
were
observed
at
24
mg/
kg/
day.
In
two
studies
reported
in
an
as
yet
unpublished
manuscript
(
Klinefelter
et
al,
2002a),
male
rats
treated
with
BCA
doses
ranging
from
8
to
216
mg/
kg/
day
showed
a
variety
of
adverse
effects,
including
significant
impairment
in
sperm
motility,
abnormal
sperm
morphology,
and
altered
spermiation.
In
one
of
these
studies,
fertility
was
assessed
by
in
utero
insemination
of
untreated
females
with
sperm
from
treated
males;
significantly
reduced
fertility
was
observed
at
all
doses
tested
(
8,
24,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
87
Draft,
do
not
cite
or
quote
and
72
mg/
kg/
day),
although
there
was
no
dose­
response.
The
LOAEL
for
the
Klinefelter
et
al.

(
2002a)
study
was
8
mg/
kg/
day
and
a
NOAEL
could
not
be
determined.

In
the
reproductive
and
developmental
screening
assay
conducted
by
NTP
(
1998),
BCA
treatment
at
50
mg/
kg/
day
for
30­
35
days
adversely
affected
the
ability
of
female
rats
to
conceive
and
carry
a
full
litter
to
term.
Adverse
reproductive
effects
were
most
prominent
early
in
gestation,
as
demonstrated
by
significantly
increased
pre­
implantation
losses
and
decreased
total
implants
per
litter,
and
nonsignificant
but
elevated
post­
implantation
losses
and
increased
number
of
resorptions.
The
statistical
power
of
the
screening
assay
was
seriously
limited
by
the
small
sample
sizes
and
the
low
number
of
pregnancies
in
each
dose
group.
Nonetheless,
based
on
statistically
significant
and
toxicologically
relevant
reproductive
and
developmental
end
points,
the
LOAEL
and
NOAEL
for
this
study
were
50
mg/
kg/
day
and
19
mg/
kg/
day,
respectively.
No
effects
on
male
reproductive
endpoints
(
testicular
histopathology,
epididymal
sperm
measures,

spermatid
head
counts,
sperm
morphology,
or
sperm
motility)
were
observed
in
the
NTP
(
1998)

screening
study.
It
is
unclear
why
these
results
differed
from
those
of
the
Klinefelter
et
al.
(
2002a)

study.
The
genotoxicity
database
for
BCA
is
very
limited.
Although
positive
results
have
been
reported
in
a
bacterial
mutagenicity
assay
(
NTP,
2000b),
in
vivo
studies
do
not
provide
evidence
of
a
direct
genotoxic
effect
(
Austin
et
al.,
1996;
Parrish
et
al.,
1996).
BCA
(
either
parent
or
metabolite)
induces
oxidative
stress
in
the
livers
of
orally­
treated
rodents
as
measured
by
increased
8­
OHdG;
these
findings
suggest
that
oxidative
DNA
damage
might
result
from
BCA
exposure.
However,
the
data
are
insufficient
to
comprehensively
evaluate
the
potential
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
88
Draft,
do
not
cite
or
quote
genotoxicity
of
BCA.
Similarly,
the
carcinogenic
potential
of
BCA
is
not
known.
In
a
published
abstract,
Stauber
et
al.
(
1995)
reported
that
BCA
induces
liver
tumors
in
mice,
but
there
are
no
published
reports
of
a
full
bioassay.
A
2­
year
NTP
toxicity
and
carcinogenesis
study
with
BCA
is
scheduled
to
be
conducted
in
the
near
future
(
NTP,
2000b).

Dibromoacetic
acid
The
toxicity
database
for
DBA
is
limited
and
has
been
developed
largely
to
explore
its
effects
on
the
male
reproductive
tract.
The
oral
LD
50
was
reported
to
be
1737
mg/
kg
in
male
rats
(
Linder
et
al.,
1994a).
The
liver
has
been
reported
to
be
a
systemic
target
organ
of
DBA­
induced
toxicity,
although
only
minimally­
adverse
effects
have
been
observed
in
short­
term
studies
(
Parrish
et
al.,
1996;
NTP,
1999).
Phillips
et
al.
(
2002,
published
abstract)
evaluated
the
neurobehavioral
toxicity
and
neuropathology
of
DBA
administered
in
drinking
water
to
male
and
female
Sprague­
Dawley
rats
for
6
months.
Neurotoxic
effects
included
mild
gait
abnormalities,

hyptonia,
decreased
forelimb
and
hindlimb
grip
strength,
decreased
sensorimotor
responsiveness
(
as
measured
by
responses
to
a
tail
pinch
and
click),
and
decreased
motor
activity.

Neuropathologic
examination
showed
significant
myelin
fragmentation,
axonal
swelling,
and
axonal
degeneration
in
the
white
matter
of
the
spinal
cord,
and
eosinophilic
or
faintly
basophilic,

occasionally
vacuolated
swelling,
indicative
of
degenerating
axons,
in
the
spinal
cord
gray
matter.

The
LOAEL
for
neurobehavioral
effects
was
20
mg/
kg/
day,
the
lowest
dose
tested.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
V­
89
Draft,
do
not
cite
or
quote
Linder
and
his
colleagues
have
studied
the
effects
of
DBA
on
spermatogenesis
and
the
resulting
consequences
for
male
fertility,
using
a
number
of
different
experimental
protocols,

including
a
single
high­
dose
study
(
Linder
et
al.,
1994a),
a
14­
day
study
(
Linder
et
al.,
1994b),

and
a
longer­
term
study
(
Linder
et
al.,
1995;
Linder
et
al.
1997a).
In
all
of
these
studies,
DBA
was
clearly
spermatotoxic
in
rats.
Based
on
histopathologic
changes
in
spermiation,
the
equivocal
LOAEL
for
the
14­
day
study
was
the
lowest
dose
tested,
10
mg/
kg/
day
(
Linder
et
al.,
1994b).
In
longer­
term
studies
in
which
male
Sprague­
Dawley
rats
were
exposed
to
DBA
for
up
to
79
days,

the
equivocal
LOAEL
for
histopathologic
changes
in
spermiation
was
10
mg/
kg/
day
and
the
corresponding
NOAEL
was
2
mg/
kg/
day.
The
severity
of
the
DBA­
induced
male
reproductivetract
toxicity
was
both
dose­
and
duration­
dependent.
Extensive
reproductive­
tract
histopathology
was
only
partially
reversed
in
rats
administered
250
mg/
kg/
day
by
oral
gavage
for
42
days
followed
by
a
6­
month
recovery
period,
indicating
that
structural
damage
to
the
reproductive
organs
was
permanent
under
the
conditions
of
this
dosing
regime.
In
an
abstract,

Veeramachaneni
et
al.
(
2000)
reported
that
rabbits
exposed
to
DBA
in
utero
from
GD
15
to
parturition,
during
lactation,
and
during
the
post­
weaning
period
through
24
weeks
of
age,

exhibited
reduced
sperm
fertility;
the
lowest
dose
tested,
0.97
mg/
kg/
day,
was
the
LOAEL.
In
contrast
to
these
findings,
a
reproductive­
toxicity
study
by
Vetter
et
al.
(
1998)
did
not
observe
significant
spermatotoxic
effects
in
male
Crl:
CD(
SD)
BR
rats
treated
with
single
oral
gavage
doses
of
600
or
1200
mg/
kg
DBA.
However,
mild
testes
histopathology
was
observed
in
both
dose
groups;
the
LOAEL
for
acute
reproductive
effects
was
600
mg/
kg
and
a
NOAEL
could
not
be
Drinking
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or
quote
determined.
The
reasons
for
the
differences
in
DBA­
induced
spermatotoxicity
between
the
Vetter
et
al.
(
1998)
study
and
those
of
Linder
and
his
colleagues
are
unclear.
In
the
recent
twogeneration
reproductive
toxicity
study
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002),

impaired
spermatogenesis
was
observed
in
male
rats
of
the
P
and
F1
generations
at
DBA
drinking
water
concentrations
of
250
ppm
and
above
(
equivalent
to
a
LOAEL
of
22
mg/
kg/
day
for
the
P
generation,
and
at
least
22
mg/
kg/
day
for
the
F1
generation);
abnormal
pathology
of
the
testes
and
epididymes
was
noted
in
some
males
of
the
F1
generation
at
650
ppm
(
equivalent
to
a
LOAEL
of
not
less
than
75
mg/
kg/
day).
However,
in
contrast
with
the
shorter­
term
study
that
showed
adverse
mating
performance
effects
at
250
mg/
kg/
day
and
higher
(
Linder
et
al.,
1995),
no
adverse
treatment­
related
effects
on
mating
performance,
gestation
length,
fertility,
pup
mortality
and
viability,
and
other
functional
indices
of
successful
reproductive
behavior
were
observed
at
DBA
drinking
water
concentrations
up
to
650
ppm
(
52
to
132
mg/
kg­
day).
Alternatively,
these
studies
in
combination
may
define
a
NOAEL/
LOAEL
boundary
for
functional
effects
of
DBA
on
reproduction.
The
weight­
of­
evidence
indicates
that
DBA
is
a
potent
male
reproductive­
system
toxicant
and
exerts
its
primary
effects
by
interfering
with
the
normal
processes
of
spermatogenesis;
however,
the
data
are
mixed
as
to
whether
these
effects
interfere
with
normal
reproductive
function.

In
a
study
on
DBA
effects
on
female
reproductive
capacity
(
Cummings
and
Hedge,
1998),

reproductive
outcomes
were
not
adversely
affected
in
rats
administered
oral
gavage
doses
of
up
to
250
mg/
kg/
day
DBA
on
GD
1­
8,
although
DBA
induced
a
significant
increase
in
serum
17­
 
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quote
estradiol
at
the
highest
dose
tested.
The
reproductive
toxicity
NOAEL
for
this
study
was
250
mg/
kg/
day
and
a
LOAEL
could
not
be
determined.
In
three
published
abstracts
summarizing
the
findings
of
developmental
toxicity
studies,
DBA
was
reported
to:
adversely
affect
pre­
and
postnatal
mortality,
decrease
pup
weight,
and
induce
skeletal
(
tail)
and
soft
tissue
(
kidney)

malformations
in
mice
exposed
in
utero
to
DBA
(
Narotsky
et
al.,
1996;
Narotsky
et
al.,
1997);

and
delay
the
pubertal
development
and
reduce
the
sperm
fertility
of
male
rats
exposed
in
utero,

during
lactation,
and
during
the
post­
weaning
period
(
to
PND
98)
to
DBA
in
drinking
water
(
Klinefelter
et
al.,
2000).
Delayed
parturition
was
also
observed
at
24
mg/
kg/
day
in
one
of
the
mouse
studies
(
Narotsky
et
al.,
1996)
but
the
biological
significance
of
this
finding
is
unclear.
The
LOAEL
and
NOAEL
for
soft­
tissue
kidney
defects
(
hydronephrosis)
in
the
mouse
study
(
Narotsky
et
al.,
1997)
were
100
and
50
mg/
kg/
day,
respectively.
In
the
pubertal
development
and
sperm­
fertility
study
(
Klinefelter
et
al.,
2000),
reduced
sperm
fertility
was
observed
in
all
dosed
male
offspring;
the
LOAEL
was
50
mg/
kg/
day
and
a
NOAEL
could
not
be
determined.
The
results
described
in
these
abstracts,
however,
cannot
be
fully
evaluated
until
a
complete
report
of
findings
is
published.
In
the
two­
generation
drinking
water
study
(
Chlorine
Chemistry
Council,

2001;
Christian
et
al.,
2002),
no
effects
were
observed
on
female
reproductive
function.

Treatment­
related
developmental
effects
included
a
statistically
significant
delay
in
preputial
separation
and
vaginal
patency,
and/
or
a
reduction
in
anogenital
distance
on
LD
22,
observed
in
F1
or
F2
pups
exposed
during
gestation
and
lactation
to
650
ppm
DBA.
These
findings
were
Drinking
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attributed
to
the
significant
growth
retardation
observed
in
these
animals,
which
was
secondary
to
decreased
water
consumption
(
due
to
taste
aversion)
by
both
pups
and
their
mothers.

The
immunotoxicity
of
DBA
administered
in
drinking
water
has
been
evaluated
in
four
studies
in
mice
(
NTP,
1999).
A
number
of
different
end
points
were
assessed,
including
thymus
and
spleen
weights,
number
and
type
of
spleen
cells,
macrophage
activation,
natural
killer
(
NK)

cell
activity,
and
specific
and
general
IgM
antibody­
forming
responses.
The
most
sensitive
and
reliable
measure
was
a
decrease
in
spleen
IgM
antibody­
forming
cell
responses,
representing
a
clear
decrease
in
immune
system
function,
accompanied
by
an
increase
in
the
number
of
spleen
macrophages.
The
LOAEL
and
NOAEL
for
these
endpoints
were
approximately
70
and
38
mg/
kg/
day,
respectively.
No
data
were
identified
for
the
toxicity
of
DBA
following
exposure
by
the
dermal
or
inhalation
routes.

The
weight­
of­
evidence
for
DBA
mutagenicity/
genotoxicity
indicates
that
DBA
is
mutagenic
and
genotoxic,
although
the
nature
of
the
DNA
damage
induced
by
DBA
remains
unclear.
The
potential
for
DBA
carcinogenicity
is
not
known.
In
published
abstracts,
So
and
Bull
(
1995)
reported
that
DBA
induces
aberrant
crypt
foci
in
the
colon
of
rats,
and
Stauber
et
al.

(
1995)
reported
that
DBA
induces
liver
tumors
in
mice.
However,
no
complete
reports
of
DBA
cancer
bioassays
have
been
published.
A
2­
year
NTP
toxicity
and
carcinogenicity
study
with
DBA
is
scheduled
to
be
conducted
in
the
near
future
(
NTP,
2000a).
Drinking
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or
quote
Chapter
VI.
Health
Effects
in
Humans
No
studies
were
identified
that
directly
evaluated
human­
health
effects
of
exposure
to
MBA,
BCA,
or
DBA
via
any
route.
Rather,
most
of
the
human­
health
data
for
brominated
acetic
acids
are
as
components
of
complex
mixtures
of
water­
disinfection
byproducts.
These
complex
mixtures
of
disinfection
byproducts
have
been
associated
with
increased
potential
for
bladder,

rectal,
and
colon
cancer
(
reviewed
by
Boorman
et
al.,
1999)
and
adverse
effects
on
reproduction
(
reviewed
by
Nieuwenhuijsen
et
al.,
1999).

Most
studies
of
human­
health
effects
following
exposure
to
water­
disinfectant
byproducts
have
used
total
trihalomethanes
as
the
exposure
metric,
and
the
risks
attributable
to
brominated
acetic
acids
typically
have
not
been
reported.
In
one
study
by
Klotz
and
Pyrch
(
1999),
a
population­
based
case­
control
study
was
conducted
on
the
relationship
between
drinking­
water
exposure
to
trihalomethanes,
haloacetonitriles,
and
haloacetic
acids
and
neural­
tube
defects.
The
study
included
112
cases
of
neural­
tube
defects
in
1993
and
1994
in
New
Jersey.
A
total
of
248
controls
were
selected
randomly
from
all
New
Jersey
births.
No
significant
relationship
between
total
trihalomethanes
and
neural­
tube
defects
was
observed
for
analysis
of
all
cases,
cases
restricted
to
subjects
with
known
residency
at
conception,
or
cases
restricted
to
isolated
cases
of
neural­
tube
defects.
However,
a
statistically
significant
difference
between
cases
and
controls
was
observed
when
cases
were
restricted
to
subjects
with
known
residency
at
conception
and
to
cases
with
isolated
neural­
tube
defects.
Based
on
this
more
stringent
case
definition,
a
prevalence
odds
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ratio
(
POR)
of
2.1
was
reported
(
95%
confidence
interval,
1.1
­
4.0)
for
the
highest
tertile
of
trihalomethane
exposure.
However,
only
a
slight
non­
statistically
significant
excess
risk
(
POR
1.2,

95%
confidence
interval
0.5­
2.6)
was
found
for
cases
when
analyzed
based
on
total
haloaceticacid
tertiles.
The
specific
haloacetic
acids
that
were
measured
as
part
of
the
total
haloacetic
acidexposure
estimate
were
not
specified.
Based
on
the
results
of
the
study,
the
authors
concluded
that
the
haloacetic
acids
did
not
exhibit
a
clear
association
with
neural­
tube
defects.
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or
quote
VII.
Mechanisms
of
Toxicity
A.
Mechanisms
of
Noncancer
Toxicity
Little
is
known
about
the
molecular
mechanisms
of
toxicity
of
the
brominated
acetic
acids.

It
has
not
been
determined
conclusively
whether
the
parent
compound
or
a
metabolite
is
the
toxic
moiety;
and
there
are
clear
differences
in
the
potency
and
spectrum
of
effects
induced
by
MBA,

BCA,
and
DBA.
For
example,
MBA
is
more
acutely
toxic
than
DBA,
but,
unlike
DBA,
is
not
spermatotoxic
(
Linder
et
al.,
1994a).
The
available
data
on
the
mechanisms
of
toxicity
of
the
brominated
acetic
acids
for
selected
endpoints
are
described
here.

One
proposed
cellular
basis
for
the
toxicity
of
MBA
is
through
direct
alkylation
of
sulfhydryl
and
amino
groups
via
its
ability
to
inhibit
a
number
of
mammalian
enzymes
in
in
vitro
studies.
However,
the
data
are
only
suggestive,
due
to
the
use
of
high
MBA
concentrations
and
purified
proteins,
and
in
vitro
test
systems.
Incubation
of
purified
human
thioredoxin
reductase
with
MBA
at
pH
6.5
inhibited
enzyme
activity
by
>
99%
(
Gorlatov
et
al.,
1998).
Although
the
concentration
of
MBA
used
in
the
reaction
was
not
presented,
the
MBA
concentration
can
be
estimated
to
be
0.7
mM,
based
on
reaction
volumes
and
moles
of
compound
used.
Similar
incubations
at
pH
6.5
and
pH
8
led
to
carboxymethylation
of
specific
selonocysteine
residues
of
thioredoxin
reductase
(
Gorlatov
et
al.,
1998),
suggesting
that
MBA
can
inhibit
enzyme
activity
through
alkylation
of
sulfhydryl
groups.
MBA
reactions
with
amino
groups
in
proteins
have
also
Drinking
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or
quote
been
suggested
as
a
mechanism
for
enzyme
inhibition
under
certain
physiological
conditions.
For
example,
Ito
et
al.
(
1994)
reported
that
100
mM
MBA
nearly
completely
inhibited
purified
human
urinary
DNase
I
activity
and
resulted
in
modifications
of
critical
histidine
residues.
Shapiro
et
al.

(
1988)
reported
that
30
mM
MBA
resulted
in
carboxymethylation
of
several
histidine
sites
in
purified
human
angiogenin
and
inactivated
the
protein.
Whitney
(
1970)
reported
the
inhibition
of
human
carbonic
anhydrase
B
following
incubation
of
human
erythrocytes
with
5
mM
MBA.
While
these
data
show
that
MBA
can
alkylate
cellular
proteins
and
disrupt
their
normal
function
in
vitro,

the
enzyme­
inhibition
studies
were
carried
out
primarily
for
the
purpose
of
identifying
critical
amino­
acid
residues
for
normal
protein
function,
not
for
determining
the
mechanism
of
action
of
MBA
toxicity.
The
threshold
concentration
for
MBA­
induced
enzyme
inhibition
was
not
reported,

and
the
relevance
of
these
findings
to
in
vivo
toxicity
is
not
clear.
In
vitro
studies
lack
a
number
of
biological
characteristics
that
can
modulate
toxicologic
responses
in
the
intact
organism,
including
hepatic
metabolism,
toxicokinetics,
and
the
presence
of
additional
protein
systems.
Further,
the
concentrations
of
MBA
used
in
these
studies
are
similar
to
or
higher
than
the
high
doses
used
in
animal
studies,
and
may
not
be
directly
comparable
to
low­
dose
environmental
exposures.
In
addition,
critical
high­
affinity
enzyme
targets
that
lead
to
the
observed
toxic
effects
of
MBA
have
not
been
identified,
although
the
alkylation
of
DNA
in
cells
in
tissue
culture
was
reported
to
induce
DNA
damage
(
Stratton
et
al.,
1981).
Taken
together,
these
data
demonstrate
the
potential
for
MBA
to
adversely
impact
cellular
macromolecules,
but
whether
this
ability
is
responsible
for
MBA
toxicity
has
not
been
clearly
shown.
Drinking
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not
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or
quote
Several
in
vivo
animal
studies
have
demonstrated
that
MBA,
BCA,
and
DBA
are
developmental
toxicants
(
Randall
et
al.,
1991
(
abstract);
NTP,
1998;
Narotsky
et
al.,
1996,
1997
(
abstracts);
Klinefelter,
2000c
(
abstract)),
although
the
spectrum
of
adverse
developmental
effects,
the
associated
toxic
potencies,
and
the
critical
periods
for
gestational
exposure
differ
significantly
among
these
three
compounds.
These
developmental
toxicity
studies
are
described
in
detail
in
Chapter
V.

The
results
of
several
mouse
whole­
embryo
testing
studies
provide
mechanistic
support
for
the
potential
for
developmental
toxicity
of
the
brominated
acetic
acids
in
vivo
and
suggest
possible
mechanisms
of
embryotoxicity.
However,
in
vitro
studies
such
as
whole
embryo
culture
(
WEC)
have
limited
utility
for
predicting
either
the
spectrum
of
adverse
developmental
effects
or
the
associated
toxic
potencies
in
intact
organisms.
In
addition
to
maternal
influences
in
the
whole
animal
during
gestation
and
lactation,
potentially
adverse
developmental
responses
observed
in
vitro
can
be
modified
by
hepatic
metabolism,
toxicokinetics,
the
activity
of
additional
protein
systems,
and
other
physiologic
and
biochemical
processes.
Further,
the
chemical
concentrations
required
to
induce
developmental
effects
in
in
vitro
experimental
systems
such
as
WEC
are
usually
much
higher
than
low­
dose
environmental
exposures.
Thus,
these
in
vitro
data
are
hypothesis­
generating
only,
and
must
be
supplemented
by
mechanistic
data
from
studies
conducted
in
vivo.
To
date,
the
data
from
in
vivo
and
in
vitro
developmental­
toxicity
studies
are
limited
and
do
not
provide
significant
information
on
possible
or
likely
mechanisms
of
Drinking
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Document
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or
quote
developmental
toxicity
for
brominated
acetic
acids,
particularly
mechanisms
which
might
explain
observed
differences
in
in
vivo
toxicity
among
brominated
acetic
acid
compounds.

In
vitro
studies
with
mouse
whole­
embryo
culture
(
WEC)
have
demonstrated
that
MBA,

DBA,
and
BCA
have
the
potential
to
induce
developmental
toxicity,
including
skeletal
(
e.
g.,

neural­
tube
defects,
pharyngeal­
arch
defects)
and
soft­
tissue
(
e.
g.,
cardiac
and
eye
defects)

malformations
(
Hunter
et
al.,
1996,
1999
(
abstract)).
Ward
et
al.
(
1997,
1998)
studied
the
effects
of
BCA
and
DBA
on
protein
kinase
C
(
PKC)
activity
in
mouse
WEC
as
a
possible
mechanism
of
developmental
toxicity
(
PKC
is
a
signal
transduction
enzyme
that
controls
the
activity
of
a
variety
of
proteins
involved
in
cell
growth
and
differentiation
via
phosphorylation).

Both
BCA
and
DBA,
in
the
concentration
range
of
0.3
­
3
mM,
inhibited
purified
rat­
brain
PKC
in
a
dose­
dependent
manner.
These
compounds
also
inhibited
PKC
activity
in
homogenates
of
GD­
9
embryos.
A
follow­
up
study
was
conducted
to
evaluate
the
relationship
between
BCA
and
DBA's
ability
to
inhibit
PKC
and
their
observed
embryotoxicity
(
Ward
et
al.,
2000).
Groups
of
6­
12
whole
CD­
1
mouse
embryos
(
early
somite­
stage
conceptuses)
were
cultured
for
up
to
24
hours
with
300
µ
M
DBA,
300
µ
M
BCA,
Bis
I
(
a
specific
PKC
inhibitor
with
previously
defined
embryotoxic
effects
(
Ward
et
al.,
1998),
staurosporine
(
a
potent,
but
non­
specific
PKC
inhibitor
known
to
interact
with
the
cell
cycle)
or
Bis
V
(
negative
control).
These
concentrations
of
BCA
and
DBA
were
chosen
to
induce
embryotoxicity
in
nearly
all
embryos
as
evidenced
by
morphological
abnormalities,
primarily
neural­
tube
defects
(
data
shown
for
DBA
only),
but
not
embryolethality.
Neither
BCA
nor
DBA
disrupted
the
cell
cycle.
However,
flow
cytometry
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
5
Draft,
do
not
cite
or
quote
revealed
the
accumulation
of
sub­
G1
events
(
indicative
of
apoptosis)
with
BCA
and
staurosporine,
but
not
DBA,
Bis
I
or
Bis
V.
For
BCA,
sub­
G1
events
were
particularly
pronounced
in
the
head
region
but
not
in
the
heart.
Although
sub­
G1
events
in
the
head
region
were
also
increased
by
DBA
treatment
(
2­
to
3­
fold
increase),
this
increase
was
not
statistically
significant.
Thus,
BCA
and
staurosporine,
but
not
Bis
I
or
DBA,
induced
apoptosis.
These
mixed
results
for
the
specific
PKC­
inhibitor
Bis
I
and
nonspecific
PKC­
inhibitor
staurosporine
make
it
unclear
whether
the
ability
of
BCA
to
inhibit
PKC
is
related
to
the
induced
apoptotic
response.

However,
because
the
two
inhibitors
have
differing
PKC­
isoform
specificities,
a
direct
role
of
PKC
inhibition
cannot
be
ruled
out.
The
study
authors
suggested
that
other
possible
mechanisms
of
dysmorphogenesis
may
include
kinase­
mediated
disruption
of
signal­
transduction
pathways
in
the
neurulation­
stage
embryo.

Hunter
et
al.
(
1999),
in
a
published
abstract,
evaluated
the
ability
of
known
haloacetic
acid
metabolites
to
induce
dysmorphogenesis
in
the
mouse
WEC
system.
The
potency
of
glycolate,

glyoxylate,
and
oxalate
were
tested.
Glycolate
induced
a
low
incidence
of
neural­
tube
defects
(
NTDs)
at
1000
µ
M,
while
no
effects
were
induced
at
this
concentration
for
glyoxylate
or
oxalate.
For
all
three
compounds,
the
severity
of
effects
increased
with
increasing
concentration.

The
concentrations
of
MBA,
DBA,
and
BCA
at
which
dysmorphogenesis
was
observed
in
the
same
test
system
were
much
lower
than
those
for
identified
metabolites.
This
result
suggests
that
the
developmental
toxicity
of
the
brominated
acetic
acids
is
not
due
to
the
metabolites
glycolate,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
6
Draft,
do
not
cite
or
quote
glyoxylate,
or
oxalate.
However,
other
as
yet
unidentified
intermediate
metabolites
may
be
implicated
in
brominated
acetic
acid
toxicity.

Andrews
et
al.
(
1999a),
in
a
published
abstract,
extended
the
use
of
whole
embryo
culture
studies
by
evaluating
the
potential
for
developmental
effects
of
BCA
and
DBA
in
rat­
embryo
cultures,
as
compared
with
mouse­
embryo
cultures
reported
previously
by
other
investigators
(
Hunter
et
al.,
1996;
Ward
et
al.,
1996,
1997).
Results
for
DBA
were
comparable
with
those
from
the
mouse­
WEC
studies
and
the
toxic
potencies
of
DBA
and
BCA
were
similar.
In
a
follow
up
study,
Andrews
et.
al.
(
1999b,
abstract)
reported
on
the
potential
for
embryotoxicity
of
mixtures
of
DCA,
DBA,
and
BCA
in
rat­
WEC.
The
experimental
results
for
the
mixtures
were
adequately
predicted
(
data
were
not
shown)
by
dose­
additivity,
as
proposed
by
a
quantitative
structure­
activity
relationship
(
QSAR)
model
developed
by
Richard
and
Hunter
(
1996;
described
below).
The
effects
on
dysmorphogenesis
were
similar
to
those
observed
in
single­
compound
WEC
(
Hunter
et
al.
1996).

The
potential
for
developmental
toxicity
among
haloacetic
compounds,
including
the
mono­,
di­,
and
tri­
substituted
fluoro­,
chloro­,
and
bromoacetic
acids
was
also
studied
using
WEC
by
Hunter
et
al.
(
1996),
who
were
mainly
interested
in
determining
if
structure­
activity
relationships
could
be
established
relating
the
type
and
degree
of
halogen
substitution
to
the
severity
and
spectrum
of
possible
developmental
effects.
For
both
MBA
and
DBA,
malformations
were
increased
at
sublethal
doses
and
the
spectrum
of
effects
was
similar,
but
MBA
was
significantly
more
potent
than
DBA.
Neither
MBA
nor
DBA
reduced
the
pH
of
the
culture
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
7
Draft,
do
not
cite
or
quote
medium,
precluding
this
as
the
mechanism
responsible
for
the
observed
developmental
toxicity.

Overall,
the
effects
of
most
haloacetic
acids
were
qualitatively
similar
and
the
ranking
of
toxic
potency
was
monobromo
>
monochloro
>
dibromo
>
trichloro,
and
tribromo
>
acetate
>
dichloro
>
trifluoro
>
difluoro.
Using
these
data,
Richard
and
Hunter
(
1996)
developed
a
QSAR
model
in
order
to
test
predictions
regarding
the
toxic
potency
of
haloacetic
acids,
and
offer
insight
into
the
mechanism(
s)
of
the
developmental
toxicity
of
this
class
of
compounds.
The
potencies
predicted
by
this
model
were
compared
with
the
potencies
determined
experimentally.
Experimentally,
the
potencies
of
the
monohaloacetic
acids
increased
with
halogen
size
(
iodo
>
bromo
>
chloro
>

fluoro),
and
the
model
was
able
to
correctly
predict
this
trend,
although
slightly
overestimating
chloroacetic
acid
potency,
and
slightly
underestimating
those
of
fluoroacetic
acid
and
bromoacetic
acid.
Experimentally,
and
as
predicted
by
the
model,
the
same
trend
held
for
the
three
dihaloacetic
acids
(
dibromo
>
dichloro
>
difluoro),
although
the
model
predicted
more
similar
potencies
of
the
difluoro­
and
the
dichloro­
compounds
than
were
seen
experimentally.
However,
the
model
was
unable
to
accurately
predict
the
toxic
potency
of
trihaloacetic
acids,
overestimating
tribromo­
and
underestimating
trichloro­
and
trifluoro­
acetic
acid
potencies.
Richard
and
Hunter
(
1996)
used
this
model
to
predict
the
developmental
toxicity
potencies
of
several
untested
haloacetic
acids,

including
BCA.
The
predicted
potency
of
BCA
was
similar
to
that
of
DBA,
and
the
relative
potencies
of
the
brominated
acetic
acids
were
MBA>
BCA

DBA.
The
results
of
the
study
demonstrated
an
increase
in
teratogenic
potency
with
increasing
pKa
values.
Since
pKa
increased
with
decreased
degree
of
halogenation,
this
relationship
accounted
for
increasing
potency
with
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
8
Draft,
do
not
cite
or
quote
decreasing
degree
of
halogenation.
The
authors
hypothesized
that
the
fit
of
the
data
supported
a
common
mechanism
of
action
for
haloacetic
acids,
with
differing
potencies
engendered
by
the
type
and
degree
of
halogen
substitution.
However,
insufficient
data
are
available
to
confirm
this
hypothesis.

The
most
well­
studied
noncancer
endpoint
of
concern
for
brominated
acetic
acids
is
male
reproductive
toxicity.
Some
evidence
also
suggests
that
liver,
kidney,
and
immune
system
toxicity
can
occur.
With
the
exception
of
effects
of
DBA
on
spermatogenesis,
the
data
are
limited
and
only
tentative
conclusions
regarding
mechanisms
of
toxicity
can
be
made.
Unifying
ideas
on
mechanisms
of
toxicity
across
the
class
of
brominated
acetic
acids
will
be
discussed
below
for
liver,
kidney,
and
reproductive
effects,
respectively.

Effects
of
oral
dosing
on
the
liver
have
been
reported
for
BCA
(
Parrish
et
al.,
1996;
NTP,

1998)
and
DBA
(
Parrish
et
al.,
1996;
NTP,
1999;
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002).
However,
in
most
cases,
minimal
evaluations
were
conducted,
and
observed
effects
were
limited
to
increased
liver
weight
and
marginal
histopathological
changes
including
cytoplasmic
vacuolization.
Both
increased
liver
weight
and
cytoplasmic
vacuolization
are
consistent
with
liver­
glycogen
accumulation
(
NTP,
1998),
a
phenomenon
that
occurs
with
dichloroacetic
acid
(
DCA)
(
Kato­
Weinstein
et
al.,
1998)

The
induction
of
lipid
peroxidation
and
oxidative
DNA
damage
in
the
livers
of
mice
treated
with
BCA
or
DBA,
in
the
absence
of
peroxisome
proliferation
(
Austin
et
al.,
1996;
Parrish
et
al.,
1996),
is
consistent
with
the
potential
for
liver
toxicity
of
these
compounds.
The
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
9
Draft,
do
not
cite
or
quote
metabolism
of
BCA
and
DBA
has
not
been
sufficiently
characterized
to
clearly
identify
the
intermediates
involved
in
lipid
peroxidation.
Both
compounds
are
apparently
metabolized
in
a
manner
similar
to
DCA,
a
weak
peroxisome
proliferator
(
De
Angelo
et
al,
1989).
Similar
reactive
intermediates
derived
from
BCA
and
DBA
might
be
responsible
for
their
ability
to
induce
lipid
peroxidation.

Recent
studies
have
demonstrated
the
metabolism
of
both
BCA
and
DBA,
as
well
as
that
of
DCA,
is
mediated
by
GST­
Zeta
(
Tong
et
al.,
1998a.
Cornett
et
al.
(
1999)
proposed
that
DCAinduced
toxicity
results
from
the
inhibition
of
GST­
Zeta,
which
is
also
known
as
maleylacetoacetate
isomerase,
an
enzyme
involved
in
tyrosine
catabolism.
Cornett
et
al.
(
1999)

found
that
DCA
exposure
increased
the
urinary
excretion
of
maleylacetone,
a
reactive
metabolite
of
tyrosine
catabolism.
Based
on
these
data,
the
authors
suggested
that
increases
in
reactive
tyrosine
metabolites
might
contribute
to
adverse
effects
induced
by
DCA.
BCA
and
DBA
also
inhibit
GST­
Zeta
activity
(
Anderson
et
al.,
1999),
suggesting
that
perturbation
of
tyrosine
catabolism
might
also
be
involved
in
the
toxicity
induced
by
brominated
acetic
acids.
The
formation
of
reactive
intermediates
or
oxidative
stress
responses
(
such
as
lipid
peroxidation)

either
by
BCA
or
DBA
directly
or
through
tyrosine
metabolites
may
be
particularly
important
in
the
potential
for
tumorigenicity
of
these
compounds.
Stauber
et
al.
(
1995)
has
reported
in
an
abstract
that
BCA
and
DBA
induced
liver
tumors
in
mice;
however,
a
complete
presentation
of
this
study
has
not
been
published
and,
thus,
these
findings
cannot
be
evaluated.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
10
Draft,
do
not
cite
or
quote
Another
potential
target
for
brominated
acetic
acids
is
the
kidney.
BCA
treatment
increased
renal
tubular
dilatation/
degeneration
in
female
rats,
but
these
changes
were
not
statistically
significant
(
NTP,
1998).
In
males
in
the
same
study,
no
treatment­
related
changes
in
kidney
weight
or
labeling
index
were
detected
and
histopathology
was
not
reported.
NTP
(
1999)

reported
increased
kidney
weight
following
oral
dosing
with
DBA
in
female
mice.
One
potential
mechanism
hypothesized
for
the
observed
renal
effects
might
be
metabolism
of
brominated
acetic
acids
to
oxalic
acid,
a
demonstrated
kidney
toxicant
that
causes
tubule
damage
by
forming
oxalate
crystals
(
Kennedy
et
al.,
1993;
Webster
et
al.,
2000),
but
this
hypothesis
has
not
been
directly
tested.
In
addition,
the
nature
and
extent
of
kidney
toxicity
induced
by
BCA
is
not
well
characterized
in
the
currently
available
data.
Further,
in
pharmacokinetic
studies
with
DCA,
the
chlorinated
analog
of
BCA,
only
2­
4%
of
the
administered
dose
was
recovered
as
urinary
oxalate
following
high­
dose
gavage
exposures
(
James
et
al.,
1998);
thus,
it
is
questionable
whether
sufficient
oxalate
would
be
formed
during
the
metabolism
of
brominated
acetic
acids
to
induce
kidney
toxicity
by
this
mechanism.

The
major
area
of
emphasis
for
toxicity
studies
for
the
brominated
acetic
acids,
has
been
on
potential
reproductive
effects.
Convincing
evidence
of
adverse
spermatogenic
effects
and
decreased
male
fertility
is
available
for
DBA
(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b;
Linder
et
al.,
1995;
Linder
et.
al.,
1997a;
Klinefelter
et
al.,
2000;
Veeramachaneni
et
al.,
2000;
;
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002)
and
BCA
(
Klinefelter
et
al.,
2002a).
In
the
single
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
11
Draft,
do
not
cite
or
quote
study
identified,
MBA
had
no
effect
on
spermatogenesis,
under
treatment
conditions
similar
to
those
that
yielded
positive
indications
of
spermatotoxicity
for
DBA
(
Linder
at
al.,
1994a).

Several
potential
mechanisms
of
male
reproductive
toxicity
have
been
explored.
Linder
et
al.
(
1994a)
identified
potential
targets
for
the
effects
of
DBA,
based
on
the
spectrum
and
timing
of
effects
on
different
stages
of
sperm
development
2,
14,
and
28
days
after
administration
of
a
single,
high
dose
of
DBA.
The
study
authors
suggested
that
DBA
induced
a
sequence
of
two
developmental
changes
in
epididymal
sperm:
abnormal
head
development
in
Step
10
or
earlier
spermatids,
and
abnormal
flagellar
development
as
the
spermatids
passed
through
the
cauda.
This
early
paper
also
noted
that
the
retention
of
Step
19
spermatids
(
the
most
common
effect
of
DBA
on
sperm
development
occurring
at
the
lowest
doses)
is
an
effect
observed
following
treatments
that
alter
hormone
status.
The
observation
that
DBA
reduced
circulating­
testosterone
levels
is
consistent
with
the
observed
effects
on
Step
19
spermatids
(
Linder
et
al.,
1994a).
Another
potential
target
for
DBA
might
be
Sertoli
cells,
because
the
presence
of
testicular
debris
suggested
to
the
study
authors
that
disruption
of
the
endocytic
activity
of
these
cells
might
be
occurring.
The
hypothesis
that
Sertoli
cells
are
a
target
for
DBA­
induced
spermatotoxicity
was
supported
by
the
results
of
a
later
multiple­
dosing
study
by
this
same
group
of
investigators
(
Linder
et
al.,
1997a).
In
this
study,
the
authors
reported
that
the
spectrum
of
late
spermatid
dysmorphogenesis
and
the
formation
of
atypical
residual
bodies
was
consistent
with
the
disruption
of
normal
Sertoli­
cell
function.
The
appearance
of
large
intraepithelial
vacuoles
in
the
Sertoli­
cell
cytoplasm
was
cited
as
additional
evidence
for
damage
to
these
cells.
The
authors
further
noted
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
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OW/
OST/
HECD
VII­
12
Draft,
do
not
cite
or
quote
that
the
disruption
of
normal
Sertoli­
cell
function
could
be
explained
by
DBA­
induced
damage
to
the
cell
cytoskeleton,
many
proteins
of
which
play
a
direct
role
in
the
developmental
functions
of
Sertoli
cells
(
Linder
et
al.,
1997a).
In
light
of
the
observed
dose­
severity
response,
with
low
DBA
doses
causing
retention
of
Step
19
spermatids
and
higher
DBA
doses
causing
overt
changes
in
sperm
morphology,
it
is
possible
that
there
are
multiple
targets
for
DBA­
induced
toxicity
in
the
male
reproductive
tract
and
that
toxicity
might
be
induced
by
more
than
one
mechanism.

Although
the
cellular
mechanisms
of
brominated
acetic
acid
spermatotoxicity
have
not
been
clearly
identified,
a
hypothesis
consistent
with
the
existing
data
is
that
the
disruption
of
normal
Sertoli­
cell
function
is
due
to
alkylation
of
critical
cellular
proteins.
There
are
multiple
pathways
by
which
DBA
might
alter
the
function
of
key
Sertoli­
cell
proteins.
For
example,
DBA
itself,
reactive
DBA
metabolites,
metabolites
formed
by
the
disruption
of
tyrosine
catabolism,

and/
or
reactive
oxygen
species
induced
secondary
to
DBA
treatment
might
be
involved
in
amino
or
sulfhydryl
group
modifications
of
critical
cellular
proteins.
Alternatively,
these
same
moieties
might
be
directly
cytotoxic,
as
suggested
by
the
observed
Sertoli­
cell
morphology
changes
following
DBA
treatment
(
Linder
et
al.,
1997a).

More
recent
work
has
tentatively
identified
one
of
the
key
protein
targets
of
brominated
acetic
acid
spermatotoxicity.
As
part
of
a
study
of
BCA
spermatotoxicity
and
fertility
assessment,

Klinefelter
et
al
(
2002a)
analyzed
120
sperm
proteins
extracted
from
male
Sprague­
Dawley
rats
treated
with
DBA
for
14
days.
A
significant
reduction
in
two
of
these
proteins,
SP22
and
SP9,

was
observed
following
treatment,
and
the
shape
of
the
dose­
response
curve
for
SP22
mirrored
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
13
Draft,
do
not
cite
or
quote
that
of
reduced
male
fertility.
The
study
authors
concluded
that
BCA,
like
DBA,
is
capable
of
perturbing
spermatogenesis
and
fertility,
and
that
SP22
appears
to
be
useful
as
a
sperm
biomarker
of
fertility.
Additional
studies
have
suggested
that
SP22
represents
a
protein
found
within
the
cytoplasm
of
round
spermatids
that
migrates
to
the
plasma
membrane
overlying
the
equatorial
segment
of
the
sperm
head
later
on
in
the
spermatogenic
process
(
Jeffrey,
1999),
and
appears
to
play
an
important
role
in
the
fertilization
process
(
Klinefelter
et
al.,
2002b).
Additionally,
the
nuclear
form
of
SP22
may
be
a
positive
regulator
of
the
androgen
receptor
(
Takahashi
et
al.,

2001),
and
it
has
been
postulated
that
haloacetic
acids
acting
on
SP22
and
other
sperm
proteins
may
indirectly
compromise
androgen­
dependent
maintenance
of
spermatogenesis
(
Klinefelter
et
al.,
2002b).

The
results
of
several
studies
have
suggested
that
DBA
may
act
as
a
reproductive
toxicant
by
interfering
with
steroidogenesis.
In
the
Linder
et
al.
(
1994a)
study,
alterations
in
sperm
parameters
at
high
doses
were
accompanied
by
a
sharp
attenuation
of
serum
testosterone
levels,

suggesting
a
steroidogenic
effect.
During
steroidogenesis,
cholesterol
is
converted
by
steroidogenic
acute
regulatory
protein
(
StAR)
and
P450
side
chain
cleavage
enzyme
(
P450
scc)
to
pregnenolone,
a
precursor
of
progesterone.
3 ­
hydroxysteroid
dehydrogenase
(
3 ­
HSD)

catalyzes
the
conversion
of
pregnenolone
to
progesterone,
which
is
then
converted
through
a
series
of
catalyzed
steps
to
testosterone.
Balchak
et
al.
(
2000)
showed
that
in
vitro
exposure
of
preovulatory
follicles,
obtained
from
immature
Sprague­
Dawley
female
rats
primed
with
a
subcutaneous
injection
of
pregnant
mare
serum
gonadotrophin,
to
50
µ
g/
mL
DBA
over
a
24­
hour
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
14
Draft,
do
not
cite
or
quote
period
blocked
release
of
progesterone
stimulated
by
human
chorionic
gonadotrophin
(
hCG)

without
having
a
comparable
effect
on
estradiol.
Using
a
similar
protocol,
Goldman
and
Murr
(
2002)
conducted
a
series
of
experiments
to
establish
a
dose­
response
for
the
effects
of
DBA
on
progesterone
secretion
and
to
identify
the
site(
s)
of
action
along
the
initial
segment
of
the
steroidogenic
pathway.
Progesterone
release
was
significantly
depressed
following
24­
hour
incubation
with
50
µ
g/
mL
DBA,
but
not
with
2
or
10
µ
g/
mL,
under
both
baseline
and
hCGstimulated
conditions.
The
suppression
in
progesterone
release
at
50
µ
g/
mL
was
shown
(
by
analysis
of
the
incubated
follicles)
to
be
due
to
a
DBA­
induced
reduction
in
follicular
progesterone
content.
No
effects
of
DBA
treatment
on
estradiol
secretion
were
observed.
In
a
second
experiment,
follicular
cultures
were
supplemented
with
pregnenolone
to
evaluate
any
effects
of
DBA
(
50
µ
g/
mL)
on
3 ­
HSD­
catalyzed
conversion
to
progesterone.
Under
these
conditions,
progesterone
release
was
increased
up
to
13­
fold
compared
to
that
of
untreated
controls.
Thus,
DBA
did
not
attenuate
progesterone
release
in
the
presence
of
pregnenolone.
The
increase
in
progesterone
secretions
observed
during
the
24­
hour
incubation
period
indicated
to
the
authors
that
DBA­
induced
depression
of
progesterone
release
observed
in
the
earlier
experiment
was
not
likely
to
be
due
to
effects
on
follicular
viability.
Determination
of
follicular
progesterone
levels
showed
that
the
progesterone
content
was
significantly
elevated
about
3­
3.5­

fold
under
both
the
hCG­
stimulated
and
non­
stimulated
conditions,
even
though
pregnenolone
supplementation
suppressed
the
DBA
attenuation
of
progesterone
release.
These
findings
imply
an
increase
in
progesterone
synthesis,
both
in
the
presence
and
absence
of
hCG
supplementation,
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
15
Draft,
do
not
cite
or
quote
that
was
not
reflected
in
the
progesterone
release
data.
Pregnenolone
treatment
did
not
have
an
effect
on
estradiol
secretion,
something
not
reflected
in
the
release
data.
In
a
third
experiment,

follicular
cultures
were
supplemented
with
22­
R
hydroxycholesterol
(
22R­
HC).
According
to
the
study
authors,
22R­
HC
can
serve
as
a
membrane­
permeable
precursor
for
pregnenolone
synthesis,

circumventing
transport
within
the
mitochondrial
membrane
by
the
StAR
protein;
its
presence
in
DBA­
treated
follicular
cultures
allows
for
the
assessment
of
DBA
effects
on
the
activity
of
P450
cholesterol
side­
chain
cleavage
enzyme
(
Pscc).
Supplementation
with
22R­
HC
eliminated
the
DBA­
induced
attenuation
effect
on
baseline
progesterone
release,
although
the
attenuation
in
the
hCG­
stimulated
secretion
was
still
present.
The
study
authors
concluded
that
exposure
to
DBA
may
have
an
effect
on
the
StAR­
mediated
transport
of
cholesterol
within
the
mitochondrial
membrane
and
an
effect
on
receptor
or
postreceptor
events
triggered
by
hCG,
but
only
at
high
doses.

B.
Cancer
Mechanisms
As
described
in
Section
V.
D,
there
are
a
number
of
studies
on
the
genotoxicity
of
MBA,

BCA,
and
DBA.
The
data
are
inadequate
for
determining
whether
MBA
or
BCA
are
genotoxic,

but
suggest
that
DBA
is
genotoxic.
Some
comparisons
to
the
chlorinated
acetic
acids
may
also
be
useful.
However,
in
the
absence
of
data
from
completed
cancer
bioassays,
consideration
of
the
mechanism
of
carcinogenicity
for
MBA,
BCA,
or
DBA
is
premature
and
overly
speculative.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
16
Draft,
do
not
cite
or
quote
C.
Sensitive
Subpopulations
No
data
are
available
to
determine
whether
sensitive
subpopulations
exist
with
regard
to
differences
in
age
or
genetic
susceptibility
The
developmental
toxicity
of
MBA,
BCA,
and
DBA
has
been
evaluated,
to
a
limited
degree,
as
described
below.
No
multi­
generation
reproductive
study
has
been
conducted
for
MBA
or
BCA,
although
a
recent
two­
generation
study
of
DBA
did
not
find
evidence
that
the
developing
fetus
is
more
sensitive
than
adults
to
the
effects
of
DBA
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002).
No
information
on
age­
dependent
changes
in
the
expression
of
genes
encoding
enzymes
thought
to
metabolize
bromoacetic
acids
was
identified.
Further,
it
is
not
yet
clear
whether
the
parent
compound
or
one
(
or
more)

intermediate
metabolites
is
the
toxic
moiety
of
concern.
In
the
absence
of
these
data,
a
full
determination
of
fetal
or
early­
age
susceptibility
and
differences
in
sensitivity
associated
with
genetic
variability
among
individuals
cannot
be
made.

For
MBA,
data
relevant
to
potential
fetal
sensitivity
is
limited
to
a
single
developmental
study
reported
in
a
published
abstract
(
Randall
et
al.,
1991).
The
induction
of
fetal
effects
only
at
doses
that
also
affected
maternal
weight
does
not
suggest
that
the
fetus
is
more
sensitive.

However,
because
the
study
was
only
presented
in
an
abstract,
and
the
available
reports
do
not
cover
the
full
range
of
developmental
endpoints,
no
firm
conclusions
can
be
drawn
from
the
limited
database.
As
for
MBA,
the
data
for
the
potential
developmental
toxicity
of
BCA
are
limited
to
a
single
reproductive
and
developmental
toxicity­
screening
assay
(
NTP,
1998).
The
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
17
Draft,
do
not
cite
or
quote
NOAEL
for
decreased
live
fetuses/
litter
and
decreased
total
implants/
litter
was
19
mg/
kg/
day,

while
this
dose
was
also
considered
a
NOAEL
for
general
toxicity
in
adult
males
and
females.

Thus,
for
both
MBA
and
BCA,
the
data
are
limited,
but
the
available
data
do
not
support
the
hypothesis
that
fetuses
or
children
are
more
sensitive
than
adults.

The
data
for
DBA
are
also
insufficient
for
a
full
evaluation
of
fetal
and
childhood
susceptibility.
In
two
published
abstracts
(
Narotsky
et
al.,
1996,
1997),
DBA
was
reported
to
induce
developmental
toxicity
in
CD­
1
mice
at
doses
lower
than
those
which
produced
maternal
toxicity.
However,
these
data
are
preliminary
and
have
only
been
reported
in
abstract
form.
Thus,

sufficient
detail
on
the
results
is
not
available
and
age­
related
differences
in
sensitivity
cannot
be
clearly
determined.
In
addition,
the
NOAEL
for
both
maternal
and
fetal
toxicity
reported
in
these
abstracts
was
well
above
the
NOAEL
for
effects
on
spermatogenesis.

Although
DBA
is
clearly
spermatotoxic
in
mature
animals
(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b;
Linder
et
al.,
1995;
Linder
et.
al.,
1997a;
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002),
the
data
are
conflicting
as
to
whether
the
developing
male
reproductive
tract
is
particularly
sensitive.
In
a
published
abstract,
Klinefelter
et
al.
(
2000)
reported
developmental
delays
(
delayed
preputial
separation)
in
male
rats
exposed
to
DBA
in
utero
from
GD15
through
PND
98
at
doses
of
50
mg/
kg/
day
and
higher.
However,
due
to
significant
individual
variability
in
this
developmental
measure,
the
statistical
and
biological
significance
of
this
finding
is
unclear,

and
insufficient
detail
is
available
from
the
published
abstract
to
evaluate
these
results.
Klinefelter
and
his
colleagues
are
currently
conducting
a
follow
up
in
utero
exposure
study
using
lower
doses
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
18
Draft,
do
not
cite
or
quote
(
personal
communication).
However,
in
the
two­
generation
reproductive/
developmental
toxicity
study
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002),
delayed
preputial
separation
in
males
and
delayed
vaginal
patency
in
females
in
the
F1
generation
was
attributed
to
a
general
retardation
of
growth
associated
with
decreased
water
intake
and
food
consumption
and
not
to
a
direct
treatment
effect.
Veeramachaneni
et
al.
(
2000)
reported
in
an
abstract
that
exposure
of
rabbits
in
utero
from
gestation
day
15
to
24
weeks
of
age
reduced
the
fertility
of
sperm
from
treated
males.
The
lowest
dose
tested,
0.97
mg/
kg/
day
was
considered
to
be
the
LOAEL.
This
LOAEL
for
fertility
changes
was
10­
fold
lower
than
the
LOAEL
of
10
mg/
kg/
day
reported
in
Linder
et
al.
(
1997a)
for
altered
histopathology.
However,
this
difference
could
reflect
interspecies
differences
(
rabbits
versus
rats),
or
differences
in
the
duration
of
dosing
(
24
weeks
versus
79
days),
as
well
as
increased
sensitivity
of
the
developing
male
reproductive
tract.

Further,
sufficient
detail
on
this
study
is
not
available
from
the
abstract
to
adequately
assess
these
findings.
A
two­
generation
study
of
DBA
administered
in
drinking
water
to
rats
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002)
found
no
evidence
that
rats
exposed
to
DBA
in
utero
and
for
the
first
71
days
of
life
are
more
susceptible
than
adults
to
the
effects
of
DBA
on
spermatogenesis.
Therefore,
the
data
are
not
sufficient
to
determine
the
relative
sensitivity
to
DBA
of
the
developing
versus
mature
male
reproductive
tract.
Similarly,
the
data
on
the
male
reproductive
effects
of
BCA
are
insufficient
to
determine
the
relative
sensitivity
to
DBA
of
the
developing
versus
mature
male
reproductive
tract.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
19
Draft,
do
not
cite
or
quote
In
contrast
to
the
lack
of
information
on
age­
dependent
differences
in
the
activity
of
enzymes
involved
in
bromoacetic
acid
metabolism,
several
genetic
differences
in
these
enzymes
have
been
identified
that
may
engender
differences
in
susceptibility
to
bromoacetic
acids.

GSTZeta
has
been
shown
to
convert
BCA
and
DBA
to
glyoxylate.
Blackburn
et
al.
(
2000)
reported
on
human
polymorphism
of
GST­
Zeta,
in
which
three
polymorphic
forms
of
the
gene,
GST*
A,

GST*
B,
and
GST*
C,
were
identified.
Based
on
in
vitro
experiments
with
purified
proteins
encoded
by
these
three
forms
of
GST­
Zeta,
GST*
A
had
a
3.6­
fold
higher
activity
toward
DCA
than
the
other
two
human
forms.
The
functional
consequences
of
the
polymorphism
of
GST­
Zeta
cannot
be
verified
in
the
absence
of
studies
of
DBA
or
BCA
metabolism
in
humans
polymorphic
for
GST­
Zeta.
Further,
it
is
unclear
what
consequences
the
observed
polymorphism
would
have
on
DBA­
or
BCA­
induced
toxicity
in
humans
because
it
is
not
known
whether
the
parent
compound
or
one
(
or
more)
of
its
metabolites
is
the
toxic
moiety.
A
similar
analysis
for
MBA
metabolism
is
not
possible
because
the
enzymes
involved
in
MBA
metabolism
have
not
yet
been
identified.

As
noted
above,
DBA
and
BCA
induce
liver
effects
consistent
with
glycogen
accumulation.
DCA,
the
chlorinated
analog
of
DBA,
has
been
shown
to
increase
hepatic
glycogen
accumulation
(
Kato­
Weinstein
et
al.,
1998)
and
these
authors
have
suggested
that
prolonged
glycogen
accumulation
can
become
irreversible,
resulting
in
liver
injury.
The
enzymatic
basis
for
increased
hepatic
glycogen
accumulation
remains
unclear.
However,
it
is
possible
that
individuals
with
glycogen­
storage
disease
(
an
inherited
deficiency
or
alteration
in
any
one
of
the
enzymes
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
20
Draft,
do
not
cite
or
quote
involved
in
glycogen
degradation)
represent
another
group
that
may
be
more
susceptible
to
DBA
or
BCA
toxicity.

Genetic
deficiencies
in
glyoxylate­
metabolism
enzymes,
including
alanine:
glyoxylate
aminotransaminase
(
AGT)
and
D­
glycerate
dehydrogenase
have
been
shown
to
be
responsible
for
primary
hyperoxaluria
type
I
and
type
II,
respectively.
These
disorders
result
in
systemic
oxalate
overload
and
induce
subsequent
kidney
toxicity
(
Webster
et
al.,
2000).
Since
bromoacetic
acid
metabolism
could
contribute
to
the
total
oxalate
load,
these
individuals
might
have
increased
susceptibility
for
kidney
toxicity.

No
quantitative
evaluation
has
been
conducted
on
the
health
impact
of
environmental
exposures
for
individuals
harboring
polymorphisms
in
genes
related
to
glycogen
storage,

antioxidant
response,
or
oxalate
synthesis.
In
each
of
these
cases,
a
significant
background
load
of
the
stressor
may
be
present;
thus,
the
excess
risk
associated
with
low
doses
of
brominated
acetic
acids
is
not
clear,
and
the
data
are
insufficient
to
determine
whether
any
of
these
groups
constitute
sensitive
populations.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
21
Draft,
do
not
cite
or
quote
D.
Interactions
No
studies
were
identified
that
evaluated
interactions
between
brominated
acetic
acids
and
chemicals
other
than
water­
disinfection
byproducts.
The
only
endpoint
for
which
mixtures
of
haloacetic
acids
have
been
evaluated
experimentally
is
developmental
toxicity
in
the
wholeembryo
culture
system.
As
described
above,
the
results
of
Andrews
et
al.
(
1999b)
in
this
in
vitro
test
system
support
the
dose­
additivity
suggested
by
Hunter
et
al.
(
1996),
but
these
findings
have
limited
utility
for
prediction
of
interactions
in
intact
animals.
Although
QSAR
by
Richard
and
Hunter
(
1996)
predicts
that
haloacetic
acids
would
be
similar
in
their
mechanisms
of
action
for
developmental
toxicity
and,
thus,
have
the
potential
for
additivity,
in
vivo
rodent
developmental
toxicity
studies
have
demonstrated
marked
differences
among
MBA,
BCA,
and
DBA
in
the
spectrum
of
induced
toxic
effects,
the
chemical
potencies
associated
with
these
effects,
and
the
critical
periods
for
gestational
exposures.
Additionally,
differences
in
the
genotoxic/
mutagenic
potential
among
MBA,
BCA,
and
DBA
suggest
that
these
compounds
might
exert
at
least
some
of
their
toxic
effects
via
distinct
mechanistic
pathways.

E.
Summary
One
proposed
cellular
basis
for
the
toxicity
of
MBA
is
its
ability
to
inhibit
enzyme
activity
through
direct
alkylation
of
sulfhydryl
and
amino
groups.
This
hypothesis
is
supported
by
in
vitro
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
22
Draft,
do
not
cite
or
quote
studies
using
purified
human
enzymes
(
Gorlatov
et
al.,
1998;
Ito
et
al.,
1994;
Shapiro
et
al.,
1988;

Whitney,
1970)
and
some
evidence
for
DNA
alkylation
(
Stratton
et
al,
1981),
but
a
direct
relationship
between
such
reactions
with
cellular
macromolecules
in
vivo
and
the
observed
toxic
effects
of
MBA
has
not
been
established,
and
there
are
several
limitations
in
extrapolating
from
the
in
vitro
data.

DBA
and
BCA
have
been
associated
with
liver,
kidney,
and
reproductive
and
developmental
toxicity
in
a
variety
of
toxicity
studies.
In
short­
term
studies,
both
BCA
(
Parrish
et
al.,
1996;
NTP,
1998)
and
DBA
(
Parrish
et
al.,
1996;
NTP,
1999)
induce
changes
in
liver
weight
and/
or
mild
histopathologic
alterations,
indicating
the
potential
for
liver
injury
with
increasing
dose
and/
or
exposure
duration.
Potential
mechanisms
for
the
induction
of
adverse
liver
effects
include
glycogen
accumulation
(
Kato­
Weinstein
et
al.,
1998),
perturbations
of
carbohydrate
homeostasis
(
Bull
et
al,
2000),
or
toxicity
due
to
the
formation
of
reactive
metabolites
from
haloacetic
acid
or
tyrosine­
metabolism
pathways
(
Austin
et
al.,
1996;
Parrish
et
al.,
1996;

Stacpoole
et
al.,
1998;
Cornett
et
al.,
1999).
The
kidney
may
also
be
a
target
for
brominated
acetic
acids
(
NTP,
1998;
NTP,
1999).
This
might
reflect
direct
toxicity
related
to
the
formation
of
reactive
metabolites
as
described
above
for
liver
toxicity,
or
may
reflect
toxicity
secondary
to
oxalate
formation
(
Kennedy
et
al.,
1993;
Webster
et
al.,
2000).

The
major
area
of
emphasis
for
toxicity
studies
for
the
brominated
acetic
acids,

particularly
for
DBA,
has
been
on
potential
reproductive
effects.
MBA
did
not
induce
spermatotoxicity
in
the
one
available
study
(
Linder
et
al.,
1994a).
DBA
induced
effects
on
sperm
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
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OW/
OST/
HECD
VII­
23
Draft,
do
not
cite
or
quote
development
at
doses
below
those
causing
overt
toxicity,
and
spermatotoxicity
at
doses
causing
overt
systemic
toxicity
(
Linder
et
al.,
1994a,
1997a).
Although
no
evidence
of
BCA­
induced
spermatotoxicity
was
found
in
the
NTP
(
1998)
reproductive
and
developmental
toxicity
screening
asssay,
Luft
et
al.
(
2000)
reported
in
an
abstract
that
BCA
decreased
male
fertility,
and
Klinefelter
et
al.
(
2002a)
demonstrated
that
BCA
impaired
sperm
quality
and
also
reduced
male
fertility.

These
data
suggest
that
both
DBA
and
BCA
are
male
reproductive
toxicants.
One
suggested
target
for
the
spermatotoxicity
is
the
Sertoli
cells
(
Linder
et
al.,
1997a).
Although
the
cellular
mechanisms
of
brominated
acetic
acid
spermatotoxicity
have
not
been
identified,
the
modification
of
key
proteins
necessary
for
Sertoli­
cell
function
or
direct
cytotoxicity
by
DBA
or
reactive
metabolites
might
be
involved.
The
results
of
several
studies
have
suggested
that
DBA
at
high
doses
may
act
as
a
reproductive
toxicant
by
interfering
with
the
early
stages
of
steroidogenesis
(
Balchak
et
al.,
2000;
Goldman
and
Murr,
2002),
possibly
by
altering
the
StAR­
mediated
transport
of
cholesterol
within
the
mitochondrial
membrane
and
thereby
affecting
the
synthesis
of
pregnenolone.
Brominated
acetic
acids
may
also
interfere
with
the
process
of
spermatogenesis.

Two
sperm
proteins,
SP22
and
SP9,
were
significantly
decreased
following
14­
day
treatment
of
male
rats
with
BCA,
and
the
shape
of
the
dose­
response
curve
for
SP22
mirrored
that
of
reduced
male
fertility
observed
in
these
animals
(
Klinefelter
et
al.,
2002a).
SP22
is
a
protein
that
appears
to
play
an
important
role
in
the
fertilization
process
(
Klinefelter
et
al.,
2002b),
possibly
by
regulating
the
androgen
receptor
(
Takahashi
et
al.,
2001),
and
it
has
been
suggested
that
Drinking
Water
Criteria
Document
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Acetic
Acids
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HECD
VII­
24
Draft,
do
not
cite
or
quote
haloacetic
acids
acting
on
SP22
and
other
sperm
proteins
may
indirectly
compromise
androgendependent
maintenance
of
spermatogenesis
(
Klinefelter
et
al.,
2002b).

All
three
brominated
acetic
acids
have
been
reported
to
induce
developmental
effects
(
Randall
et
al.,
1991;
NTP,
1998;
Narotsky
et
al.,
1996;
Narotsky
et
al.,
1997),
although
the
spectrum
of
developmental
endpoints
observed
does
not
necessarily
suggest
a
common
mechanism
of
action
for
the
in
vivo
studies.
Results
of
whole­
embryo
testing
were
consistent
with
a
common
mechanism
of
action,
because
a
QSAR
model
adequately
described
the
potency
of
a
mono­
and
di­
halogenated
series
of
haloacetic
acids
(
Hunter
et
al.,
1996;
Richard
and
Hunter,

1996);
further
testing
of
haloacetic­
acid
mixtures
in
whole­
embryo
culture
was
consistent
with
the
QSAR
model
predictions
(
Andrews
et
al.,
1999b).
Brominated
acetic
acids
also
induced
dysmorphogenesis
at
doses
lower
than
their
known
metabolites
in
the
whole­
embryo
testing
system
(
Hunter
et
al.,
1999),
suggesting
that
the
parent
compound
or
unidentified
metabolites
upstream
of
glyoxylate
are
responsible.
Ward
et
al.
(
2000)
proposed
a
potential
role
of
apoptosis
induction
in
developmental
toxicity
of
brominated
acetic
acids,
based
on
results
in
a
wholeembryo
culture
system.
It
should
be
noted,
however,
that
although
the
findings
from
wholeembryo
culture
systems
indicate
the
potential
for
developmental
toxicity
of
brominated
acetic
acids,
these
in
vitro
results
are
limited
in
their
utility
to
predict
both
the
spectrum
of
effects
and
the
toxic
potencies
of
these
compounds
in
in
vivo
animal
systems
due
to
the
modulating
influences
of
a
variety
of
other
physiologic
and
biochemical
processes
in
intact
organisms.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VII­
25
Draft,
do
not
cite
or
quote
There
are
no
data
available
for
identifying
susceptible
populations.
A
two
generation
reproductive
study
exists
for
DBA
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002),
but
no
multi­
generation
studies
have
been
conducted
for
MBA
or
BCA.
In
addition,
no
data
on
agedependent
changes
in
the
expression
of
genes
involved
in
brominated
acetic
acids
were
found.

Based
on
the
results
of
in
vivo
developmental
toxicity
studies,
DBA,
but
not
MBA
or
BCA,

induced
fetal
toxicity
at
doses
lower
than
those
associated
with
maternal
effects,
suggesting
that,

at
least
for
DBA,
the
fetus
might
be
susceptible.
These
results
were
only
published
in
abstracts
and,
thus,
complete
study
reports
are
not
available
to
evaluate
these
findings.
However,
these
preliminary
studies
found
fetal
and
maternal
effects
only
at
doses
well
above
those
causing
effects
on
sperm
and
male
reproduction,
indicating
that
protection
against
the
latter
effect
will
also
provide
adequate
protection
to
children
and
fetuses.

There
are
also
limited
data
on
potential
susceptible
populations
based
on
genetic
differences.
Blackburn
et
al.
(
2000)
characterized
human
polymorphisms
in
GST­
Zeta,
which
metabolizes
DBA
and
BCA
to
glyoxylate.
However,
in
the
absence
of
data
on
whether
the
parent
compound
or
a
metabolite
is
the
active
moiety,
the
functional
consequences
of
this
polymorphism
with
regard
to
brominated
acetic
acid
toxicity
are
not
clear.
Individuals
having
underlying
defects
in
glycogen
storage
may
be
susceptible
to
liver
effects
of
brominated
acetic
acids,
and
individuals
lacking
certain
enzymes
of
glyoxylate
metabolism
may
be
at
risk
for
BCA
or
DBA­
induced
kidney
toxicity.
If
the
formation
of
reactive
oxygen
or
lipid
intermediates
is
responsible
for
the
toxicity
of
brominated
acetic
acids,
then
deficits
in
the
activity
of
anti­
oxidant
enzymes
could
also
represent
a
Drinking
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Document
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OST/
HECD
VII­
26
Draft,
do
not
cite
or
quote
source
of
increased
susceptibility.
All
of
these
possibilities
remain
speculative,
and
none
has
been
tested
directly
in
in
vivo
studies.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
1
Draft,
do
not
cite
or
quote
Chapter
VIII.
Quantification
of
Toxicological
Effects
The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.
Chemicals
that
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur.
Carcinogens
are
assumed
to
act
without
a
threshold
unless
there
are
data
elucidating
a
nonmutagenic
mode
of
action
and
demonstrating
a
threshold
for
the
precursor
events
that
commit
a
cell
to
an
abnormal
tumorigenic
response.

A.
Introduction
to
Methods
A.
1.
Quantification
of
Noncarcinogenic
Effects
A.
1.1.
Reference
Dose
In
quantification
of
noncarcinogenic
effects,
a
Reference
Dose
(
RfD)
(
formerly
called
the
Acceptable
Daily
Intake
(
ADI))
is
calculated
(
U.
S.
EPA,
2001).
The
RfD
is
"
an
estimate
(
with
uncertainty
spanning
approximately
an
order
of
magnitude)
of
a
daily
exposure
to
the
human
population
(
including
sensitive
subgroups)
that
is
likely
to
be
without
appreciable
risk
of
deleterious
effects
over
a
lifetime"
(
U.
S.
EPA,
1993).
The
RfD
is
derived
from
a
no­
observed­
Drinking
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Draft,
do
not
cite
or
quote
adverse­
effect
level
(
NOAEL),
lowest­
observed­
adverse­
effect
level
(
LOAEL),
or
a
NOAEL
surrogate
such
as
a
benchmark
dose
identified
from
a
subchronic
or
chronic
study,
and
divided
by
a
composite
uncertainty
factor(
s).
The
RfD
is
calculated
as
follows:

RfD
=
NOAEL
or
LOAEL
UF
×
MF
where:

NOAEL
=
No­
observed­
adverse­
effect
level
from
a
high­
quality
toxicological
study
of
an
appropriate
duration
LOAEL
=
Lowest­
observed­
adverse­
effect
level
from
a
high­
quality
toxicological
study
of
an
appropriate
duration.
In
situations
where
there
is
no
NOAEL
for
a
contaminant
but
there
is
a
LOAEL,
the
LOAEL
can
be
used
for
the
RfD
calculation
with
the
inclusion
of
an
additional
uncertainty
factor.

UF
=
Uncertainty
factor
chosen
according
to
EPA/
NAS
guidelines
MF
=
Modifying
factor
Selection
of
the
uncertainty
factor
to
be
employed
in
calculation
of
the
RfD
is
based
on
professional
judgment
while
considering
the
entire
database
of
toxicological
effects
for
the
chemical.
To
ensure
that
uncertainty
factors
are
selected
and
applied
in
a
consistent
manner,
the
Office
of
Water
(
OW)
employs
a
modification
to
the
guidelines
proposed
by
the
National
Academy
of
Sciences
(
NAS,
1977,
1980).
According
to
the
EPA
approach
(
U.
S.
EPA,
1993),

uncertainty
is
broken
down
into
its
components,
and
each
component
of
uncertainty
is
given
a
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
3
Draft,
do
not
cite
or
quote
quantitative
rating.
The
total
uncertainty
factor
is
the
product
of
the
component
uncertainties.
The
individual
components
of
the
uncertainty
are
as
follows:

UF
H
A
factor
of
1,
3,
or
10
used
when
extrapolating
from
valid
data
in
studies
using
long­
term
exposure
to
average
healthy
humans.
This
factor
is
intended
to
account
for
the
variation
in
sensitivity
(
intraspecies
variation)
among
the
members
of
the
human
population.

UF
A
An
additional
factor
of
1,
3,
or
10
used
when
extrapolating
from
valid
results
of
long­
term
studies
on
experimental
animals
when
results
of
studies
of
human
exposure
are
not
available
or
are
inadequate.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
animal
data
to
humans
(
interspecies
variation).

UF
S
An
additional
factor
of
1,
3,
or
10
used
when
extrapolating
from
less­
thanchronic
results
on
experimental
animals
when
there
are
no
useful
long­
term
human
data.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
less­
than­
chronic
NOAELs
to
chronic
NOAELs.

UF
L
An
additional
factor
of
1,
3,
or
10
used
when
deriving
an
RfD
from
a
LOAEL,
instead
of
a
NOAEL.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
LOAELs
to
NOAELs.

UF
D
An
additional
factor
of
1,
3,
or
10
used
when
deriving
an
RfD
from
an
"
incomplete"
database.
This
factor
is
meant
to
account
for
the
inability
of
any
single
type
of
study
to
consider
all
toxic
endpoints.
The
intermediate
factor
of
3
(
approximately
½
log
10
unit,
i.
e.,
the
square
root
of
10)
is
often
used
when
there
is
a
single
data
gap
exclusive
of
chronic
data.
It
is
often
designated
as
UF
D.

On
occasion,
EPA
also
uses
a
modifying
factor
in
the
determination
of
the
RfD.
A
modifying
factor
is
an
additional
uncertainty
factor
that
is
greater
than
zero
and
less
than
or
equal
to
10.
The
magnitude
of
the
MF
depends
upon
the
professional
assessment
of
scientific
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
4
Draft,
do
not
cite
or
quote
uncertainties
of
the
study
and
database
not
explicitly
treated
above
(
e.
g.,
the
number
of
species
tested).
The
default
value
for
the
MF
is
1.

In
establishing
the
UF
or
MF,
it
is
recognized
that
professional
scientific
judgment
must
be
used.
The
total
product
of
the
uncertainty
factors
and
modifying
factor
should
not
exceed
3000.
If
the
assignment
of
uncertainty
results
in
a
UF/
MF
product
that
exceeds
3000,
then
the
database
does
not
support
development
of
an
RfD.
The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.

Unless
otherwise
specified,
chemicals
which
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.

A.
1.2.
Drinking
Water
Equivalent
Level
The
drinking
water
equivalent
(
DWEL)
is
calculated
from
the
RfD.
The
DWEL
represents
a
drinking­
water­
specific
lifetime
exposure
at
which
adverse,
noncarcinogenic
health
effects
are
not
anticipated
to
occur.
The
DWEL
assumes
100%
exposure
from
drinking
water.
The
DWEL
provides
the
noncarcinogenic
health­
effects
basis
for
establishing
a
drinking­
water
standard.
For
ingestion
data,
the
DWEL
is
derived
as
follows:
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
5
Draft,
do
not
cite
or
quote
DWEL
=
(
RfD)
×
BW
WI
where:

BW
=
70­
kg
adult
body
weight
WI
=
Drinking
water
intake
(
2
L/
day)

A.
1.3.
Health
Advisory
Values
In
addition
to
the
RfD
and
the
DWEL,
EPA
calculates
Health
Advisory
(
HA)
values
for
noncancer
effects.
HAs
are
determined
for
lifetime
exposures
as
well
as
for
exposures
of
shorter
duration
(
1­
day,
10­
day,
and
longer­
term).
The
shorter­
duration
HA
values
are
used
as
informal
guidance
to
municipalities
and
other
organizations
when
emergency
spills
or
contamination
situations
occur.
The
lifetime
HA
becomes
the
MCLG
for
a
chemical
that
is
not
a
carcinogen.

The
shorter­
term
HAs
are
calculated
using
an
equation
similar
to
the
RfD
and
DWEL;

however,
the
NOAELs
or
LOAELs
are
derived
from
acute
or
subchronic
studies
and
identify
a
sensitive,
noncarcinogenic
endpoint
of
toxicity.
The
HAs
are
derived
as
follows:

HA
=
NOAEL
or
LOAEL
×
BW
UF
×
WI
where:

NOAEL
or
LOAEL
=
No­
or
lowest­
observed­
adverse­
effect­
level
in
mg/
kg
bw/
day
BW
=
Assumed
body
weight
of
a
child
(
10
kg)
or
an
adult
(
70
kg)
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
6
Draft,
do
not
cite
or
quote
UF
=
Uncertainty
factor,
in
accordance
with
EPA
or
NAS/
OW
guidelines
WI
=
Assumed
daily
water
intake
of
a
child
(
1
L/
day)
or
an
adult
(
2
L/
day)

Using
the
above
equation,
the
following
drinking­
water
HAs
are
developed
for
noncarcinogenic
effects:

°
1­
day
HA
for
a
10­
kg
child
ingesting
1
L
water
per
day.

°
10­
day
HA
for
a
10­
kg
child
ingesting
1
L
water
per
day.

°
Longer­
term
HA
for
a
10­
kg
child
ingesting
1
L
water
per
day.

°
Longer­
term
HA
for
a
70­
kg
adult
ingesting
2
L
water
per
day.

Each
of
these
shorter­
term
HA
values
assumes
that
the
total
exposure
to
the
contaminant
comes
from
drinking
water.

The
lifetime
HA
is
calculated
from
the
DWEL,
and
takes
into
account
exposure
from
sources
other
than
drinking
water.
It
is
calculated
using
the
following
equation:

Lifetime
HA
=
DWEL
×
RSC
where:

DWEL=
Drinking
water
equivalent
level
RSC
=
Relative
source
contribution.
The
fraction
of
the
total
exposure
allocated
to
drinking
water.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
7
Draft,
do
not
cite
or
quote
A.
2
Quantification
of
Carcinogenic
Effects
Under
the
new
U.
S.
EPA
(
1999)
draft
cancer
risk
assessment
guidelines,
the
U.
S.
EPA
assesses
the
carcinogenic
potential
of
a
chemical
compound
in
a
narrative
characterization,
and
uses
one
of
the
following
five
standard
descriptors
to
express
the
conclusion
regarding
the
weight
of
evidence
for
carcinogenic
hazard
potential:

°
Carcinogenic
to
Humans
°
Likely
to
be
Carcinogenic
to
Humans
°
Suggestive
Evidence
of
Carcinogenic
Potential
°
Inadequate
Information
to
Assess
Carcinogenic
Potential
°
Not
Likely
to
be
Carcinogenic
to
Humans
Each
standard
descriptor
is
presented
only
in
the
context
of
a
chemical­
specific,
weight­

ofevidence
narrative.
Additionally,
more
than
one
conclusion
may
be
reached
for
an
agent
(
e.
g.,
an
agent
is
"
likely
to
carcinogenic"
by
inhalation
exposure
and
"
not
likely
to
be
carcinogenic"
by
oral
exposure.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
8
Draft,
do
not
cite
or
quote
If
toxicological
evidence
leads
to
the
classification
of
the
contaminant
as
a
genotoxic,

probable
or
possible
human
carcinogen,
mathematical
models
are
used
to
calculate
the
estimated
excess
cancer
risk
associated
with
ingestion
of
the
contaminant
in
drinking
water.
The
data
used
in
these
estimates
usually
come
from
lifetime­
exposure
studies
in
animals.
In
order
to
predict
the
risk
for
humans
from
animal
data,
animal
doses
must
be
converted
to
equivalent
human
doses.

This
conversion
includes
correction
for
noncontinuous
exposure,
less­
than­
lifetime
studies
and
differences
in
size.
It
is
assumed
that
the
average
adult
human­
body
weight
is
70
kg
and
that
the
average
water
consumption
of
an
adult
human
is
two
liters
of
water
per
day.

For
contaminants
with
a
carcinogenic
potential,
chemical
levels
are
correlated
with
a
carcinogenic­
risk
estimate
by
employing
a
cancer
potency
(
unit
risk)
value
together
with
the
assumption
for
lifetime
exposure
via
ingestion
of
water.
Under
the
1986
Carcinogen
Risk
Assessment
Guidelines,
the
cancer
unit
risk
was
usually
derived
from
a
linearized
multistage
model
with
a
95%
upper
confidence
limit
providing
a
low­
dose
estimate;
that
is,
the
true
risk
to
humans,
while
not
identifiable,
is
not
likely
to
exceed
the
upper­
limit
estimate
and,
in
fact,
may
be
lower.
Excess
cancer­
risk
estimates
may
also
be
calculated
using
other
models
such
as
the
one­
hit,

Weibull,
logit
and
probit
models.
There
is
little
basis
in
the
current
understanding
of
the
biological
mechanisms
involved
in
cancer
to
suggest
that
any
one
of
these
models
is
able
to
predict
risk
more
accurately
than
any
of
the
others.
Because
each
model
is
based
upon
differing
assumptions,
the
estimates
that
are
derived
for
each
model
can
differ
by
several
orders
of
magnitude.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
2
A"
point
of
departure"
(
POD)
marks
the
beginning
of
extrapolation
to
lower
doses.
The
POD
is
an
estimated
dose
(
expressed
inhuman­
equivalent
terms)
near
the
lower
end
of
the
observed
range,
without
significant
extrapolation
to
lower
doses
(
U.
S.
EPA,
2003).

EPA/
OW/
OST/
HECD
VIII­
9
Draft,
do
not
cite
or
quote
Under
the
new
U.
S.
EPA
(
1999)
draft
cancer
risk
assessment
guidelines,
dose­
response
assessment
is
performed
in
two
steps:
assessment
of
observed
experimental
data
to
derive
a
point
of
departure
(
POD)
2,
followed
by
extrapolation
to
lower
exposure
to
the
extent
that
is
necessary
for
environmental
exposures
of
interest.
Extrapolation
is
based
on
extension
of
a
biologicallybased
model
if
supported
by
substantial
data.
Otherwise,
default
approaches
can
be
applied
that
are
consistent
with
current
understanding
of
mode(
s)
of
action
of
the
agent.
These
approaches
may
assume
either
linearity
or
nonlinearity
of
the
dose­
response
relationship,
or
both.
The
linear
approach
is
used
when
there
is
an
absence
of
sufficient
information
on
modes
of
action
or
the
mode
of
action
information
indicates
that
the
dose­
response
curve
a
low
dose
is
or
is
expected
to
be
linear.
A
range
of
models
may
be
used
for
the
linear
approach.
A
default
approach
for
nonlinearity
can
be
to
use
a
reference
dose
or
a
reference
concentration
(
U.
S.
EPA,
1999).

The
scientific
data
base
used
to
calculate
and
support
the
setting
of
cancer­
risk
rates
has
an
inherent
uncertainty
due
to
the
systematic
and
random
errors
in
scientific
measurement.
In
most
cases,
only
studies
using
experimental
animals
have
been
performed.
Thus,
there
is
uncertainty
when
the
data
are
extrapolated
to
humans.
When
developing
cancer­
risk
rates,
several
other
areas
of
uncertainty
exist,
such
as
the
incomplete
knowledge
concerning
the
health
effects
of
contaminants
in
drinking
water,
the
impact
of
the
experimental
animal's
age,
sex,
and
species,
the
nature
of
the
target
organ
system(
s)
examined
and
the
actual
rate
of
exposure
of
the
internal
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
10
Draft,
do
not
cite
or
quote
targets
in
experimental
animals
or
humans.
Dose­
response
data
usually
are
available
only
for
high
levels
of
exposure,
not
for
the
lower
levels
of
exposure
at
which
a
standard
may
be
set.
When
there
is
exposure
to
more
than
one
contaminant,
additional
uncertainty
results
from
a
lack
of
information
about
possible
synergistic
or
antagonistic
effects.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
11
Draft,
do
not
cite
or
quote
B.
Noncarcinogenic
Effects
B.
1
Monobromoacetic
acid
Table
VIII­
1
summarizes
the
available
studies
on
the
oral
toxicity
of
MBA.

Table
VIII­
1.
Summary
of
Oral
Studies
of
MBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
(
mg/
kg/
day)

Linder
et
al.,
1994a
Sprague
Dawley
Rat
(
male)
Oral
Gavage
in
water
Acute
Single
dose;
100
to
200
mg/
kg
Lethality,
clinical
observation
­
LD50
177
mg/
kg
Doses
not
specified.

Linder
et
al.,
1994a
Sprague
Dawley
Rat
(
male)
Oral
Gavage
in
water
Acute
Single
dose;
0,
100
mg/
kg
Sperm
analysis,
reproductive­
tract
histopathology
100
­
None
Randall
et
al.,
1991
Long­
Evans
Rat
(
female)
Oral
Gavage
in
water
Gestation
day
6­
15;
0,
25,
50,
100
mg/
kg/
day
Decreased
maternal
weight
gain,
decreased
live­
fetus
size,
increased
incidence
of
softtissue
malformations
50
100
Published
abstract
does
not
provide
adequate
details
for
definitive
review.

Linder
et
al.,
1994a
Sprague
Dawley
Rat
(
male)
Oral
Gavage
in
water
14
day;
0,
25
mg/
kg/
day
Sperm
analysis,
reproductive­
tract
histopathology
25
­
Only
listed
endpoints
were
evaluated
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
12
Draft,
do
not
cite
or
quote
B.
1.1
One­
Day
Health
Advisory
for
MBA
The
oral
toxicity
data
for
MBA
are
very
limited.
Linder
et
al.
(
1994a)
reported
an
LD
50
of
177
mg/
kg
in
Sprague­
Dawley
rats.
Clinical
signs
included
excess
drinking,
hypomobility,
labored
breathing,
and
diarrhea.
However,
LD
50
studies
are
not
suitable
for
the
development
of
one­
day
health
advisories.
The
only
other
single­
dose
study
was
reported
by
these
same
authors,
where
male
Sprague­
Dawley
rats
were
given
0
or
100
mg/
kg/
day
MBA
and
were
evaluated
for
evidence
of
spermatotoxicity.
No
other
endpoints
were
evaluated.
The
single
dose
tested
was
the
approximate
LD
01
and
the
NOAEL
for
male
reproductive
effects
in
this
study.
Due
to
the
absence
of
an
observed
effect,
the
use
of
a
single
dose
precluding
assessment
of
the
dose­
response,
and
the
limited
endpoints
evaluated,
this
study
is
not
sufficient
for
derivation
of
a
One­
day
health
advisory.

B.
1.2
Ten­
Day
Health
Advisory
for
MBA
Two
studies
of
appropriate
duration
were
identified
for
derivation
of
a
Ten­
day
health
advisory.
However,
limited
details
or
inadequate
study
designs
preclude
their
use.
In
a
published
abstract,
Randall
et
al.
(
1991)
evaluated
the
developmental
toxicity
of
MBA
in
female
Long­

Evans
rats
dosed
with
0,
25,
50,
or
100
mg/
kg/
day
MBA
on
gestation
days
6­
15.
The
authors
reported
decreased
maternal
weight­
gain,
decreased
live­
fetus
size,
and
increased
incidence
of
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
13
Draft,
do
not
cite
or
quote
soft­
tissue
malformations
at
the
highest
dose.
Based
on
these
data,
the
NOAEL
would
be
50
mg/
kg/
day
and
the
LOAEL
for
maternal
and
developmental
effects
would
be
100
mg/
kg/
day.

However,
this
study
is
available
only
as
a
published
abstract
that
has
not
undergone
scientific
peer
review.
Thus,
the
data
are
to
be
viewed
as
preliminary,
and
the
results
of
this
study
are
not
sufficient
for
derivation
of
a
health
advisory.
A
single
published
study
of
adequate
duration
was
identified
for
MBA.
In
this
study,
no
effects
on
the
male
reproductive
tract
were
seen
in
male
rats
treated
with
0
or
25
mg/
kg/
day
MBA
by
gavage
in
water
for
14
days
(
Linder
et
al.,
1994a);
no
LOAEL
was
identified.
The
absence
of
an
identified
effect,
the
use
of
a
single
dose
level
precluding
assessment
of
the
dose­
response,
and
the
limited
array
of
endpoints
examined
preclude
the
use
of
this
study
as
the
basis
for
derivation
of
the
Ten­
day
health
advisory.
In
light
of
the
qualitative
differences
in
toxicity
between
MBA
and
DBA,
it
is
not
appropriate
to
assume
that
male
reproductive
toxicity
is
the
most
sensitive
endpoint
for
MBA,
and
it
is,
therefore,
not
appropriate
to
consider
this
free­
standing
NOAEL
for
a
limited
number
of
endpoints
as
sufficiently
protective
for
general
toxicity,
nor
to
use
it
for
derivation
of
a
health
advisory.

B.
1.3
Longer­
Term
Health
Advisory
for
MBA
There
are
no
studies
of
suitable
duration
for
derivation
of
a
Longer­
Term
Health
Advisory
for
MBA.
Developmental­
toxicity
data,
such
as
that
reported
in
an
abstract
by
Randall
et
al.

(
1991),
are
appropriate
for
the
derivation
of
Longer­
Term
health
advisories
only
if
systemic­
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
14
Draft,
do
not
cite
or
quote
toxicity
studies
of
adequate
duration
have
been
conducted
and
show
that
developmental
toxicity
is
the
most
sensitive
endpoint.
In
the
absence
of
such
systemic­
toxicity
studies
for
MBA,
no
Longer­

Term
Health
Advisory
can
be
derived.

B.
1.4
Reference
Dose
and
Drinking
Water
Equivalent
Level
for
MBA
There
are
no
studies
(
subchronic
or
chronic
toxicity
studies
that
evaluate
a
range
of
systemic
endpoints)
suitable
for
derivation
of
an
RfD
for
MBA.
As
for
the
Longer­
Term
Health
Advisory,
developmental­
toxicity
data
are
not
appropriate
for
the
derivation
of
an
RfD
in
the
absence
of
systemic­
toxicity
studies
of
adequate
duration.
In
addition,
the
only
available
developmental­
toxicity
study
(
Randall
et
al.,
1991)
was
published
as
an
abstract.

B.
2
Bromochloroacetic
acid
Table
VIII­
2
summarizes
the
available
studies
on
the
oral
toxicity
of
BCA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
15
Draft,
do
not
cite
or
quote
Table
VIII­
2.
Summary
of
Oral
Studies
of
BCA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
(
mg/
kg/
day)

NTP,
1998
Sprague
Dawley
Rat
(
male
and
female)
Drinking
Water
14
day;
0,
3,
10,
28,
41
mg/
kg/
day
Clinical
observation,
body
weight,
body
weight
gain
41
­
None
Luft
et
al.,
2000
C57BL/
6
Mouse
(
male)
Oral
Gavage
in
water
14
day;
0,
8,
24,
72,
216
mg/
kg/
day
Decrease
in
mean
number
of
litters
per
male,
decreased
percent
of
litters
per
bred
female
24
72
Published
abstract
does
not
provide
adequate
details
for
definitive
review.

Results
of
histopathology
analysis
were
not
reported.

Klinefelter
et
al.,
(
2002a),
manuscript
Sprague­
Dawley
Rat
(
male)
Oral,
Gavage
in
water
14
day,
0,
24,
72,
216,
mg/
kg/
day
(
range­
finding
study)
Decreased
serum
hormone
levels,
sperm
abnormalities,
altered
spermiation
­
24
The
LOAEL
was
the
lowest
dose
tested.

Klinefelter
et
al.,
(
2002a),
manuscript
Sprague­
Dawley
Rat
(
male)
Oral,
Gavage
in
water
14
day,
0,
8,
24,
72
mg/
kg/
day
(
dose­
response
study)
Decreased
progressive
sperm
motility,
significantly
reduced
fertility,
decrease
in
sperm
protein
SP22.
­
8
Decreased
serum
hormone
levels
not
observed.

Decrease
in
sperm
protein
SP22
paralleled
the
reduction
in
fertility.

The
LOAEL
was
the
lowest
dose
tested.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
2.
Summary
of
Oral
Studies
of
BCA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
(
mg/
kg/
day)

EPA/
OW/
OST/
HECD
VIII­
16
Draft,
do
not
cite
or
quote
Parrish
et
al.,
1996
B6C3F1
Mouse
(
male)
Drinking
water
21
days;
0,
25,
125,
500
mg/
kg/
day
Increased
liver
weight,
increased
oxidative
DNA
damage
125
500
500
mg/
kg/
day
is
considered
a
marginal
LOAEL.

Doses
were
calculated
from
default
waterintake
estimates.

NTP,
1998
Sprague
Dawley
Rat
(
male)
Drinkin
g
water
30
or
26
days;
0,
5,
15,
39
mg/
kg/
day
Liver
weight
and
histopathology
,
sperm
quality
15
39
Marginal
LOAEL;
males
had
marginal
liver
weight
and
histopathology
changes
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
2.
Summary
of
Oral
Studies
of
BCA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
(
mg/
kg/
day)

EPA/
OW/
OST/
HECD
VIII­
17
Draft,
do
not
cite
or
quote
NTP,
1998
Sprague
Dawley
Rat
(
female)
Drinking
water
Group
A:
34
days
periconception
(
12
days
premating
up
to
5
days
cohabitation,
and
up
to
21
days
gestation);
0,
6,
19,
50
mg/
kg/
day
Group
C:
30
days
periconception
(
12
days
premating
up
to
5
days
cohabitation,
and
up
to
16
days
gestation);
0,
6,
19,
50
mg/
kg/
day
Group
B:
17
days
­
gestation
day
6
to
postnatal
day
1;
0,
10,
25,
61
mg/
kg/
day
Decreased
live
fetuses
per
litter
and
decreased
total
implants
per
litter
Decreased
live
fetuses
per
litter,
decreased
total
implants
per
litter,
kidney
histopathology
changes
Maternal
and
fetal
toxicity
19
19
61
50
50
­
Effects
on
fetuses
were
based
on
a
pooled
analysis
of
Group
A
and
Group
C
data
Kidney
histopathological
changes
are
considered
as
a
equivocal
NOAEL
at
19
mg/
kg/
day.

Increase
in
postimplantation
loss
was
not
statistically
significant
and
lacked
a
doseresponse
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
18
Draft,
do
not
cite
or
quote
B.
2.1
One­
Day
Health
Advisory
for
BCA
No
studies
of
suitable
duration
were
identified
for
derivation
of
a
One­
day
health
advisory
for
BCA.

B.
2.2.
Ten­
Day
Health
Advisory
for
BCA
Five
studies
of
suitable
duration
were
identified
for
derivation
of
a
Ten­
day
health
advisory
for
BCA.
However,
one
of
these
studies
was
available
only
as
a
published
abstract
(
Luft
et
al.,
2000)
and
two
of
the
available
studies
either
identified
only
marginal
effects
or
were
not
designed
to
evaluate
systemic
toxicity
(
Parrish
et
al.,
1996;
dose­
range
finding
study
by
NTP,

1998).
The
reproductive/
developmental
effects
in
the
remaining
two
studies
of
BCA
(
NTP,
1998;

Klinefelter
et
al.,
2002a)
were
not
considered
to
be
suitable
for
the
derivation
of
a
10­
Day
health
advisory
for
a
10­
kg
child,
because
the
study
was
conducted
on
sexually­
mature
animals.

Although
equivocal
liver
effects
(
marginal
increases
in
liver
weights
and
marginal
histopathology)

were
observed
in
the
NTP
(
1998)
study,
there
was
no
dose­
response
and
no
effect
on
hepatic
labeling
index
indicative
of
cellular
proliferation
and
regeneration.
Further,
similar
marginal
histopathology
(
i.
e.,
mild
cytoplasmic
vacuolization)
was
also
observed
in
control
animals.
A
nonstatistically
significant
increase
in
kidney­
tubule
dilatation/
degeneration
was
also
observed
in
dosed
animals,
but
it
was
unclear
whether
these
findings
were
treatment­
related.
Other
organs
and
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
19
Draft,
do
not
cite
or
quote
endpoints
were
not
evaluated
in
this
study
and
sample
sizes
were
small,
thus
limiting
the
statistical
power
of
this
experiment.
These
studies
are
discussed
in
detail
in
Chapter
V.
The
limitations
of
these
studies,
including
low
statistical
power
and
toxicity
evaluation
of
only
a
small
number
of
endpoints,
preclude
their
use
in
the
derivation
of
a
Ten­
day
health
advisory.

B.
2.3.
Longer­
Term
Health
Advisory
for
BCA
The
toxicity
database
for
BCA
is
very
limited.
There
are
no
studies
of
sufficient
duration
for
derivation
of
Longer­
term
health
advisory
for
BCA.
As
noted
for
MBA
and
in
the
previous
section
for
BCA,
the
developmental
effects
noted
in
the
NTP
(
1998)
study
are
inadequate
as
the
basis
for
the
Longer­
term
health
advisory,
in
the
absence
of
a
subchronic
study
that
adequately
evaluated
systemic
toxicity.
No
multi­
generation
reproductive
toxicity
study
has
been
conducted.

Subchronic
and
chronic
toxicity
testing
of
BCA
is
planned
or
in
progress
(
NTP,
2000b).

B.
2.4
Reference
Dose
and
Drinking
Water
Equivalent
Level
for
BCA
As
discussed
in
the
previous
section
on
the
Longer­
term
health
advisory,
the
toxicity
database
for
BCA
is
currently
limited
and
there
are
no
suitable
studies
of
appropriate
design
and
duration
to
derive
an
oral
RfD
at
this
time.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
20
Draft,
do
not
cite
or
quote
B.
3
DBA
Table
VIII­
3
summarizes
the
available
studies
on
the
oral
toxicity
of
DBA.

Table
VIII­
3.
Summary
of
Oral
Studies
of
DBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
mg/
kg/
day
General
Toxicity
Studies
(
by
duration
of
treatment)

Linder
et
al.,
1994a
Sprague­
Dawley
rat
(
male)
Oral
Gavage
in
water
Acute
single
dose;
1000
to
2000
mg/
kg
Lethality,
clinical
observation
­­
LD50
1737
mg/
kg
Doses
not
specified.

Parrish
et
al.,
1996
B6C3F1
mouse
(
male)
Drinking
water
21
days;
0,
25,
125,
500
mg/
kg/
day
Increased
liver
weight,
oxidative
DNA
damage
25
125
Doses
were
calculated
from
default
water
intake
estimates
NTP,
1999
B6C3F1
mouse
(
female)
Drinking
water
28
days;

Study
(
1);
0,
19,
39,
73,
150,
285
mg/
kg/
day
Study
(
2);
0,
20,
38,
70,
143,
280
mg/
kg/
day
Study
(
3);
0,
16,
35,
69,
134,
229
mg/
kg/
day
Study
(
4);
0,
14,
33,
68,
132,
236
mg/
kg/
day
Decreased
antibody­
forming
cell
response
38
70
Absolute
and
relative
liver
weight
were
increased
beginning
at
14
mg/
kg/
day.
This
was
not
chosen
as
the
critical
effect
in
the
absence
of
histopathology
or
clinical
chemistry
data
to
confirm
that
the
effect
was
adverse.

Phillips
et
al.,
(
2002)
F344
rat
(
male
and
female
adolescents)
Drinking
water
6
months,
0,
20,
72,
161
mg/
kg/
day
Neuromuscular
and
neurobehavioral
abnormalities
in
all
dosed
groups,
spinal
cord
neuropathology
in
mid­
and
highdose
groups
­
20
Published
abstract.
Neurotoxicity
and
neuropathology
only
end
points
examined.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
3.
Summary
of
Oral
Studies
of
DBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
mg/
kg/
day
EPA/
OW/
OST/
HECD
VIII­
21
Draft,
do
not
cite
or
quote
Reproductive
Toxicity
Studies
(
by
duration
of
treatment)

Linder
et
al.,
1994a
Sprague­
Dawley
rat
(
male)
Oral
Gavage
in
water
Acute
single
dose,
up
to
28­
day
recovery;
0
or
1250
mg/
kg
Reproductiveorgan
weight
changes,
decreased
serum
testosterone,
sperm­
quality
changes,

reproductivetract
histopathology
­­
1250
A
single
dose
level
was
used.

Vetter
et
al.,
1998
Crl:
CD
(
SD)
Br
rat
(
male)
Oral
Gavage
in
water
Acute
single
dose;
0,
600,
1200
mg/
kg
Testes
histopathology
­
600
Sperm
analysis
was
limited
to
motility
and
membrane
permeability,
with
no
adverse
effects
reported.

Cummings
and
Hedge,
1998
Holtzman
rat
(
female)
Oral
Gavage
in
water
Gestation
days
1­
8;
0,
62.5,
125,
250,
500
mg/
kg/
day
Clinical
observation
250
500
(
FEL)
Reproductive
parameters
were
not
affected
at
250
mg/
kg/
day
or
less
and
were
not
measured
in
the
high
dose
group
due
to
overt
toxicity.

Linder
et
al.,
1994b
Sprague­
Dawley
rat
(
male)
Oral
Gavage
in
water
14
daily
doses;
0,
10,
30,
90,
or
270
mg/
kg
Reproductivetract
histopathology
­­
10
None
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
3.
Summary
of
Oral
Studies
of
DBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
mg/
kg/
day
EPA/
OW/
OST/
HECD
VIII­
22
Draft,
do
not
cite
or
quote
Linder
et
al.,
1995
Sprague­
Dawley
rat
(
male)
Oral
Gavage
in
water
2,
5,
9,
16,
31,
or
42
days;
0,
250
mg/
kg/
day
Decreased
reproductive
performance
on
Day
8
 
14
mating
and
day
15
­
21
mating.
Decreased
fertility
of
male
sperm
day
16
and
31.
Altered
sperm
parameters
beginning
on
day
9.
­­
250
None
Linder
et
al.,
1995
and
Linder
et
al.,
1997a
Sprague­
Dawley
rat
(
male)
Oral
Gavage
in
water
Up
to
79
days;
0,
2,
10,
or
50
mg/
kg/
day
42
days;
250
mg/
kg/
day
Reproductivetract
histopathology
2
10
(
equivocal)
Histopathology
analysis
presented
in
Linder
et
al.,
1997
Effects
at
LOAEL
became
significant
at
31
days.

Veeramachaneni
et
al.,
2000
Dutchbelted
rabbits
(
male)
Drinking
water
Dams
gestation
days
15
through
life,
male
offspring
through
24
weeks;
0,
0.97,
5.05,
54.2
mg/
kg/
day
At
24
weeks,
fertility
of
sperm
tested
in
artificially
inseminated
does
nd
0.97
Published
abstract
does
not
provide
adequate
details
for
definitive
review.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
3.
Summary
of
Oral
Studies
of
DBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
mg/
kg/
day
EPA/
OW/
OST/
HECD
VIII­
23
Draft,
do
not
cite
or
quote
Christian
et
al.,
1999
Sprague­
Dawley
rat
(
male
and
female)
Drinking
water
Sires
from
study
day
(
SD)
1­
70:
10,
20,
36,
66
mg/
kg/
day;
Dams
from
SD
1­
15:
15,
30,
49,
82
mg/
kg/
day;
from
gestation
day
(
GD)
0­
21:
15,
39,
49,
82
mg/
kg/
day;
from
lactation
day
(
LD)
1­
29:
44,
87,
151,
212
mg/
kg/
day
Slight
nonsignificant
decrease
in
mating
performance
and
number
of
mated
pairs
at
highest
dose
tested;
reduced
body
wt
gain,
body
wt,
water
consumption,
food
intake
attributed
to
taste
aversion
of
DBAtreated
water
66
(
sires)

>
60
(
dams)

>
82
(
developmental
­
Mean
daily
water
intake
and
corresponding
mean
DBA
daily
doses
significantly
increased
in
pregnant
and
lactating
females
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
3.
Summary
of
Oral
Studies
of
DBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
mg/
kg/
day
EPA/
OW/
OST/
HECD
VIII­
24
Draft,
do
not
cite
or
quote
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
(
2002)
Sprague­
Dawley
rat
(
male
and
female)
Drinking
water
Two­
generation
reprod.
toxicity
study,
0,
50,
250,
650
ppm
in
water;
estimated
daily
doses
are:
P
males
­
SD
1­
92:
0,
4.4,
22,
52
mg/
kg/
day;
P
females
­
SD
1­
70:
0,
6,
28,
69
mg/
kg/
day;
GD
0­
21:
0,
6,
30,
76
mg/
kg/
day;
LD
1­
15:
0,
12,
56,
132;
F1
males:
0,
5­
6,
22­
30;
55­
75
mg/
kg/
day;
F1
females:
Premating:
0,
7,
32,
83
mg/
kg/
day;
GD
0­
21:
0,
6,
29,
67
mg/
kg/
day;
LD
1­
15:
0,
10,
50,
115
mg/
kg/
day
Impaired
spermatogenesis,
testicular
histopathology
in
P
and
F1
males;
no
treatmentrelated
effects
in
females
Reduced
body
wt
gain,
body
wt,
water
consumption,
food
intake
in
P,
F1,
F2
animals
and
changes
in
organ
weights
in
P
and
F1
attributed
to
taste
aversion
effects
of
DBA­
treated
water
4
(
P
males)

>
5
(
F1
males)
22
(
P
males)

>
22
(
F1
males)
Study
report
was
reviewed
by
an
independent
scientific
advisory
panel
Recently
published
in
International
Journal
of
Toxicology
21:
237­
276,
2002.

Developmental
Toxicity
Studies
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
Table
VIII­
3.
Summary
of
Oral
Studies
of
DBA
Toxicity
Reference
Species
Route
Exposure
Duration;
Doses
Endpoints
NOAEL
LOAEL
Comments
mg/
kg/
day
EPA/
OW/
OST/
HECD
VIII­
25
Draft,
do
not
cite
or
quote
Narotsky
et
al.,
1996
CD­
1
mouse
(
female)
Oral
Gavage
in
water
Gestation
days
6­
15;
0,
24,
50,
100,
200,
392,
610,
or
806
mg/
kg/
day
Increased
postnatal
mortality;
decreased
pup
weight,
tail
defects
392
610
Published
abstract
does
not
provide
adequate
details
for
definitive
review.

Maternal
toxicity
was
limited
to
decreased
maternal
motor
activity
at
the
high
dose.

Narotsky
et
al.,
1997
CD­
1
mouse
(
female)
Oral
Gavage
in
water
Gestation
days
6­
15;
0,
50,
100,
or
400
mg/
kg/
day
Fetal
malformations;
hydronephrosis
50
100
Published
abstract
does
not
provide
adequate
details
for
definitive
review.

Hydronephrosis
and
renal
agenesis
at
400
mg/
kg/
day.

B.
3.1
One­
Day
Health
Advisory
for
DBA
The
acute
oral
toxicity
data
for
DBA
are
very
limited.
Linder
et
al.
(
1994a)
reported
an
LD
50
of
1737
mg/
kg
in
Sprague­
Dawley
rats.
However,
LD
50
studies
are
not
suitable
for
the
development
of
One­
day
health
advisories.
In
the
same
paper,
Linder
et
al.
(
1994a)
assessed
the
effects
of
a
single
oral
dose
of
0
or
1250
mg/
kg/
day
on
the
male
reproductive
system.
After
dosing,
the
animals
were
followed
for
up
to
28
days.
The
single
dose
tested
was
spermatotoxic,

and
induced
severe
changes
in
sperm­
quality
parameters
and
reproductive­
tract
histopathology.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
26
Draft,
do
not
cite
or
quote
The
absence
of
dose­
response
data
limits
the
utility
of
this
study
for
health­
advisory
derivation.
In
a
similar
acute­
dosing
study
validating
a
new
test
method,
Vetter
et
al.
(
1998)
also
evaluated
the
spermatotoxicity
of
DBA
in
male
Crl:
CD(
SD)
Br
rats
given
single
oral
doses
of
0,
600,
or
1200
mg/
kg
DBA.
The
high
dose,
but
not
the
low
dose,
resulted
in
clinical
observations
of
toxicity
and
testes
histopathology,
but
no
effects
on
sperm
motility,
morphology,
or
cell­
membrane
permeability;
analysis
was
limited
to
evaluation
of
these
measures.
Due
to
the
limited
doseresponse
and
the
testing
for
a
limited
number
of
endpoints,
this
study
is
not
suitable
for
deriving
a
One­
day
health
advisory.

B.
3.2
Ten­
Day
Health
Advisory
for
DBA
A
number
of
toxicity
studies
have
been
reported
for
DBA
that
are
of
suitable
duration
for
derivation
of
a
Ten­
day
health
advisory.
These
studies
have
evaluated
reproductive­
and
developmental­
toxicity
endpoints,
as
well
as
some
indices
of
systemic
toxicity.

Two
studies
by
Narotsky
and
colleagues
(
Narotsky
et
al.,
1996;
Narotsky
et
al.,
1997)

observed
adverse
developmental
effects,
including
skeletal
and
soft­
tissue
malformations,
in
the
offspring
of
pregnant
mice
administered
DBA
by
gavage
on
GD
6­
15.
Full
study
reports
have
not
been
published,
so
these
studies
are
not
appropriate
for
the
derivation
of
a
health
advisory.

Male
fertility
and
sperm
parameters
were
also
evaluated
in
groups
of
rats
administered
0
or
250
mg/
kg/
day
DBA
by
gavage
for
2­
42
days
(
Linder
et
al.,
1995),
but
only
a
single
high
dose
was
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
27
Draft,
do
not
cite
or
quote
tested.
Linder
et
al.
(
1994b)
identified
an
equivocal
LOAEL
of
10
mg/
kg/
day
based
on
histopathological
changes
of
the
seminiferous
tubules
in
adult
male
Sprague­
Dawley
rats
administered
14
daily
gavage
doses
of
DBA;
a
NOAEL
could
not
be
determined.
In
a
study
of
female
reproductive
function
and
fetal
development,
gavage
doses
of
up
to
125
mg/
kg/
day
on
GD
1­
8
had
no
effects
on
reproductive
parameters
or
clinical
observations
of
toxicity
in
rats
(
Cummings
and
Hedge,
1998).
The
highest
dose
tested
in
this
study
(
500
mg/
kg/
day)
was
lethal
and
not
evaluated
for
reproductive
outcome.
None
of
these
reproductive
studies
are
appropriate
for
the
derivation
of
a
Ten­
day
health
advisory
for
a
10­
kg
child
because
the
findings
are
only
relevant
to
sexually­
mature
animals
and
not
to
children.

Two
other
studies
were
of
an
appropriate
duration
for
derivation
of
a
Ten­
day
health
advisory,
but
these
studies
did
not
include
evaluation
of
a
complete
array
of
systemic
endpoints.

Parrish
et
al.
(
1996)
evaluated
the
ability
of
DBA
to
induce
oxidative
DNA
damage
in
the
livers
of
mice
treated
with
DBA
in
drinking
water
for
21
days.
Increased
liver
weight
and
levels
of
8­

OHdG,
a
measure
of
oxidative
stress,
were
observed
at
125
mg/
kg/
day,
but
the
absence
of
histopathology
or
clinical
chemistry
data
makes
it
unclear
whether
the
observed
increase
in
liver
weight
was
adverse.
Other
organ
systems
and
end
points
were
not
evaluated.
In
an
immunotoxicity­
screening
assay
comprised
of
4
short­
term
studies
(
NTP,
1999),
female
mice
treated
with
DBA
in
the
drinking
water
for
28
days
exhibited
an
array
of
immunotoxic
effects;

however,
many
of
these
effects
were
inconsistent
and/
or
did
not
exhibit
a
dose­
response.
Based
on
decreased
spleen
IgM
antibody­
forming
response,
the
NOAEL
was
38
mg/
kg/
day
and
the
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
28
Draft,
do
not
cite
or
quote
LOAEL
was
70
mg/
kg/
day.
Thymus
and
spleen
weights
were
also
evaluated,
but
other
organs
and
end
points
were
not
assessed.
Therefore,
neither
of
these
studies
(
Parish
et
al.,
1996;
NTP,

1999)
are
considered
to
be
suitable
for
derivation
of
a
Ten­
day
health
advisory
in
the
absence
of
other
toxicity
studies
that
have
adequately
evaluated
systemic
toxicity.

B.
3.3
Longer­
Term
Health
Advisory
for
DBA
A
number
of
studies
that
examined
the
reproductive
or
developmental
toxicity
of
DBA
were
evaluated
for
the
potential
derivation
of
a
Longer­
term
HA.
As
described
previously,

published
abstracts
are
available
for
two
developmental­
toxicity
studies
in
mice
(
Narotsky
et
al.,

1996,
1997),
which
demonstrated
adverse
developmental
effects
including
increased
postnatal
mortality,
and
skeletal
(
tail
defects)
and
soft­
tissue
(
kidney
defects)
malformations.
A
published
abstract
for
a
neurotoxicity
study
in
rats
has
shown
that
DBA
produces
neurobehavioral
toxicity,

including
neuromuscular
abnormalities,
decreased
sensorimotor
responsiveness,
and
increased
motor
activity,
as
well
as
spinal
cord
neuropathology
indicative
of
axonal
degeneration
(
Phillips
et
al.,
2002).
However,
none
of
these
studies
have
been
published
and,
thus,
these
results
are
not
suitable
for
the
derivation
of
human­
health
advisories.

Linder
et
al.
(
1995,
1997a)
evaluated
the
spermatotoxicity
and
fertility
in
male
Sprague­

Dawley
rats
administered
daily
gavage
doses
of
DBA
of
0,
2,
10,
or
50
mg/
kg/
day
for
up
to
79
days.
In
a
companion
study,
male
rats
were
gavaged
with
either
0
or
250
mg/
kg/
day
daily
for
42
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
29
Draft,
do
not
cite
or
quote
days,
at
which
time
dosing
was
terminated
due
to
severe
overt
toxicity.
Fertility
in
the
dosed
males
was
assessed
through
day
213
by
mating
treated
males
with
untreated
females
at
different
time
periods.
Based
on
the
results
of
these
studies,
DBA
is
clearly
spermatotoxic
and
effects
on
sperm
histopathology
appear
to
be
the
most
sensitive
endpoint,
because
these
effects
are
observed
in
the
absence
of
other
reproductive
toxicity
endpoints.
Changes
in
retention
of
Step
19
spermatids
was
the
only
effect
that
occurred
at
the
lowest
dose.
This
effect
was
equivocally
noted
following
repeated
dosing
with
10
mg/
kg/
day,
but
not
2
mg/
kg/
day,
for
31
or
79
days.
However,

the
biological
significance
of
this
finding
for
a
Longer­
term
human­
health
advisory
is
unclear
because
changes
in
sperm
count,
morphology,
and
motility
were
observed
at
higher
doses
(
50
mg/
kg/
day)
than
those
associated
with
these
early
and
mild
histopathological
changes,
and
male
fertility
was
significantly
affected
only
at
250
mg/
kg/
day
(
Linder
et
al.,
1995).
At
doses
of
50
mg/
kg/
day
and
lower,
there
were
no
significant
effects
on
reproductive
outcome
as
indicated
by
a
number
of
different
measures
and
indices
of
successful
mating
behavior
were
not
significantly
altered.
These
results
are
described
in
detail
in
Section
V.
Although
it
has
been
proposed
that
the
fertility
of
rodents
may
be
less
sensitive
to
changes
in
sperm
count
than
fertility
in
humans
(
U.
S.

EPA,
1996a;
Zenick
et
al.,
1994),
human
data
are
highly
variable
and
generally
inconsistent
across
studies.
Further,
there
may
be
significant
differences
in
the
susceptibility
of
different
species
and
rodent
strains
to
DBA­
induced
reproductive
toxicity.
In
an
acute
spermatotoxicity
study
by
Vetter
et
al.
(
1998),
changes
in
sperm
motility
and
morphology
were
not
observed
in
male
Crl:
CD(
SD)
BR
rats
at
acute
gavage
doses
of
up
to
1200
mg/
kg,
whereas
Linder
et
al.
(
1994a)
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
30
Draft,
do
not
cite
or
quote
observed
significant
alterations
in
both
of
these
parameters
in
male
Sprague­
Dawley
rats
following
an
acute
gavage
dose
of
1250
mg/
kg.
A
published
abstract
by
Veeramachaneni
et
al.

(
2000)
reported
that
male
Dutch­
belted
rabbits
exposed
in
utero
from
GD
15
through
lactation
and
post­
weaning
for
24
weeks
exhibited
decreased
fertility,
as
evidenced
by
reduced
conception
in
females
artificially
inseminated
with
sperm
from
treated
animals,
at
drinking­
water
doses
as
low
as
0.97
mg/
kg/
day.
However,
a
full
report
of
this
study
has
not
been
published,
and
thus
these
results
cannot
be
comprehensively
evaluated.
It
is
also
not
known
whether
humans
would
be
less,

or
more,
sensitive
to
DBA­
induced
male
reproductive­
tract
toxicity
than
rats
or
rabbits.
No
reproductive
epidemiologic
data
on
DBA
are
available,
and
comparative
in
vitro
studies
have
not
been
conducted.

Although
the
studies
by
Linder
et
al.
(
1994a,
1994b,
1995,
1997)
adequately
characterize
the
male
reproductive
hazards
in
rats
repeatedly
administered
DBA
by
oral
gavage,
these
studies
are
not
considered
to
be
suitable
for
quantitative
dose­
response
assessment
in
the
absence
of
subchronic
and
chronic
toxicity
studies
that
have
adequately
evaluated
DBA
systemic
toxicity.
In
the
recent
two­
generation
reproductive/
developmental
toxicity
study
conducted
by
the
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002),
impaired
spermatogenesis
was
also
observed
in
male
rats
of
the
P
and
F1
generations
at
DBA
drinking
water
concentrations
of
250
ppm
and
above
(
equivalent
to
a
LOAEL
of
22
mg/
kg/
day
for
the
P
generation,
and
not
less
than
22
mg/
kg/
day
for
the
F1
generation);
abnormal
pathology
of
the
testes
and
epididymes
was
noted
in
some
males
of
the
F1
generation
at
650
ppm
(
equivalent
to
a
LOAEL
of
not
less
than
75
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
31
Draft,
do
not
cite
or
quote
mg/
kg/
day).
The
corresponding
NOAEL
was
4
mg/
kg/
day
in
the
P
generation
and
at
least
5
mg/
kg/
day
in
the
F1
generation.
However,
in
contrast
with
the
shorter­
term
study
that
showed
adverse
mating
performance
effects
at
250
mg/
kg/
day
and
higher
(
Linder
et
al.,
1995),
no
adverse
treatment­
related
effects
on
mating
performance,
gestation
length,
fertility,
pup
mortality
and
viability,
and
other
functional
indices
of
successful
reproductive
behavior
were
observed
at
DBA
drinking
water
concentrations
up
to
650
ppm
(
52
to
132
mg/
kg­
day)
(
Chlorine
Chemistry
Council,
2001;
Christian
et
al.,
2002).
Alternatively,
these
studies
in
combination
may
define
a
NOAEL/
LOAEL
boundary
for
functional
effects
of
DBA
on
reproduction.
No
treatment­
related
adverse
developmental
effects
other
than
impaired
spermatogenesis
were
noted
in
either
males
or
females
of
the
F1
and
F2
generations.
To
date,
other
developmental
toxicity
studies
have
only
been
published
in
abstract
form
and
thus
cannot
be
comprehensively
evaluated
until
full
study
reports
are
available
for
review.
Subchronic
and
chronic
toxicity
testing
of
DBA
is
planned
or
in
progress
(
NTP,
2000c).
A
number
of
additional
studies
are
currently
ongoing.
Therefore,
it
is
not
appropriate
to
develop
a
Longer­
term
health
advisory
at
this
time.

B.
3.4.
Reference
Dose
and
Drinking
Water
Equivalent
Level
for
DBA
As
discussed
in
the
previous
section
on
the
Longer­
term
health
advisory,
the
toxicity
database
for
DBA
is
currently
limited,
and
there
are
no
suitable
studies
of
appropriate
design
and
duration
to
derive
an
oral
RfD
at
this
time.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
32
Draft,
do
not
cite
or
quote
C.
Carcinogenic
Effects
No
epidemiology
or
animal
studies
were
identified
to
develop
a
quantitative
cancer­
risk
assessment
for
MBA,
BCA,
or
DBA.
No
studies
were
identified
that
directly
evaluated
the
human
carcinogenicity
of
MBA,
BCA,
or
DBA.
Rather,
most
of
the
human­
health
data
for
brominated
acetic
acids
are
as
components
of
complex
mixtures
of
water­
disinfection
byproducts.
These
complex
mixtures
of
disinfection
byproducts
have
been
associated
with
increased
potential
for
cancer
(
Boorman
et
al.,
1999),
but
brominated
acetic
acids
have
not
been
specifically
implicated.

C.
1.
Monobromoacetic
acid
No
epidemiology
or
animal
studies
were
identified
that
evaluated
the
carcinogenicity
of
MBA.
The
data
are
inadequate
for
determining
whether
MBA
is
genotoxic.
The
mutagenicity
of
MBA
might
be
metabolism­
dependent,
based
on
different
results
for
these
compound
in
the
presence
or
absence
of
microsomal
activation
in
S.
typhimurium
strain
TA100.
MBA
was
reported
to
be
mutagenic
in
three
independent
studies
in
bacteria
(
Giller
et
al.,
1997;
Kohan
et
al.,

1998;
NTP
2000a),
but
MBA
did
not
induce
DNA
repair
as
measured
by
the
SOS
chromotest
(
Giller
et
al.,
1997).
Due
to
the
limited
database,
there
is
insufficient
evidence
to
determine
the
genotoxicity
of
MBA.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
33
Draft,
do
not
cite
or
quote
Following
the
EPA's
1986
(
U.
S.
EPA,
1986)
Guidelines
for
Cancer
Risk
Assessment,

MBA
is
best
classified
as
Group
D,
"
not
classifiable
as
to
human
carcinogenicity".
This
classification
is
appropriate
because
no
data
are
available
on
human
or
animal
carcinogenicity.

Under
the
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999b),
the
data
are
"
inadequate
for
an
assessment
of
human
carcinogenic
potential"
of
MBA.

C.
2.
Bromochloroacetic
acid
No
epidemiology
studies
have
evaluated
the
carcinogenicity
of
BCA.
The
carcinogenicity
of
BCA
has
not
been
tested
in
a
full
cancer
bioassay.
However,
BCA
is
currently
slated
for
testing
(
NTP,
2000b).
The
only
carcinogenicity­
testing
data
in
animals
for
BCA
that
was
identified
was
reported
in
a
published
abstract
(
Stauber
et
al.,
1995).
The
abstract
reported
preliminary
data
suggesting
that
BCA
induces
hepatic
tumors
in
B6C3F1
mice.
However,
no
experimental
details
were
provided
in
the
brief
study
summary
and
the
full
study
report
has
not
been
published.
BCA
was
reported
as
positive
in
the
single
standard
assay
identified,
a
Salmonella
reverse­
mutation
assay
(
NTP,
2000b).
The
reports
of
Austin
et
al.
(
1996)
and
Parrish
et
al.
(
1996)
demonstrated
that
BCA
treatment
could
induce
oxidative
DNA
damage
in
the
livers
of
treated
mice.
While
these
data
are
suggestive
of
genotoxic
potential,
BCA
has
not
been
sufficiently
tested
to
make
a
determination
as
to
its
genotoxicity.
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
34
Draft,
do
not
cite
or
quote
Following
the
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment,
BCA
is
best
classified
as
Group
D,
"
not
classifiable
as
to
human
carcinogenicity"
(
U.
S.
EPA,
1986).
This
classification
is
appropriate
because
no
data
are
available
on
human
carcinogenicity
and
there
are
only
preliminary
animal
carcinogenicity
data.
Under
the
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999b),
the
data
are
"
inadequate
for
an
assessment
of
human
carcinogenic
potential"
of
BCA.

C.
3.
Dibromoacetic
acid
No
epidemiology
studies
have
evaluated
the
carcinogenicity
of
DBA.
The
carcinogenicity
of
DBA
has
not
been
tested
in
a
full
cancer
bioassay.
However,
DBA
is
currently
undergoing
testing
(
NTP,
2000c).
In
published
abstracts,
So
and
Bull
(
1995)
reported
that
DBA
induces
aberrant
crypt
foci
in
the
colon
of
rats,
and
Stauber
et
al.
(
1995)
reported
that
DBA
induces
liver
tumors
in
mice.
Experimental
details
are
not
available
for
either
of
these
studies
because
neither
has
been
published
in
peer­
reviewed
journals,
but
the
findings
of
So
and
Bull
(
1995)
might
be
of
particular
significance
since
colon
cancer
has
been
implicated
as
a
potential
cancer
site
in
humans
exposed
to
drinking­
water
disinfectant
by­
products,
including
haloacetic
acids
(
Boorman
et
al.,

1999).

Much
of
the
concern
for
the
potential
carcinogenicity
of
DBA
arises
from
the
demonstrated
high­
dose
rodent­
liver
tumorigenicity
of
its
chlorinated
analog,
dichloroacetic
acid
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
35
Draft,
do
not
cite
or
quote
(
DCA).
The
ability
of
both
compounds
to
induce
a
similar
spectrum
of
noncarcinogenic
toxic
effects
suggests
the
possibility
that
this
might
also
be
the
case
for
carcinogenic
effects.
For
example,
both
compounds
are
potent
spermatotoxicants
and
induce
a
similar
spectrum
of
effects
in
the
male
reproductive
tract
(
Linder
et
al.,
1997b).
Both
compounds
also
affect
the
liver.

Treatment
with
DBA
or
DCA
resulted
in
increased
liver
weight,
although
only
DBA
increased
the
formation
of
oxidative
DNA
damage
in
this
study
(
Parrish
et
al.,
1996).
DBA
and
DCA
also
appear
to
have
similar
kinetics
(
Schultz
et
al.,
1999),
but
insufficient
data
are
available
on
DBA
metabolism,
renal
elimination,
and
tissue
distribution
to
fully
compare
the
kinetics
of
DBA
and
DCA.
These
similarities
between
DBA
and
DCA
suggest
that,
like
DCA,
DBA
might
also
be
tumorigenic
at
high
drinking­
water
doses
administered
for
a
lifetime.
However,
the
weight­

ofevidence
for
DCA
genotoxicity
indicates
that
DCA
is
nongenotoxic
except
possibly
at
high
doses
that
also
induce
cytotoxicity.
In
contrast,
although
the
DBA
database
for
genotoxicity/

mutagenicity
is
more
limited
than
that
for
DCA,
the
weight­
of­
evidence
to
date
indicates
that
DBA
is
genotoxic
Thus,
although
both
DBA
and
DCA
exhibit
similar
toxicokinetics
and
similar
systemic
and
reproductive
toxicity,
they
may
well
differ
in
their
carcinogenic
potential
and
mode(
s)
of
carcinogenic
action.

Thus,
insufficient
data
are
available
to
assess
DBA
carcinogenic
hazard,
and
DBA
is
classified
as
Group
D,
"
not
classifiable
as
to
human
carcinogenicity"
under
the
1986
Carcinogen
Risk
Assessment
Guidelines.
Under
the
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
VIII­
36
Draft,
do
not
cite
or
quote
(
U.
S.
EPA,
1999b),
the
data
are
"
inadequate
for
an
assessment
of
human
carcinogenic
potential"
of
DBA.

D.
Summary
In
the
absence
of
a
comprehensive
toxicity
database,
no
adequate
studies
of
suitable
design
and/
or
duration
were
identified
to
serve
as
the
basis
for
any
health
advisories
for
MBA,

BCA,
or
DBA.

MBA,
BCA,
and
DBA
are
all
classified
as
"
not
classifiable
as
to
human
carcinogenicity"
under
the
1986
Carcinogen
Risk
Assessment
Guidelines
,
and
"
inadequate
for
an
assessment
of
human
carcinogenic
potential"
under
the
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999b).
Drinking
Water
Criteria
Document
for
Brominated
Acetic
Acids
EPA/
OW/
OST/
HECD
IX­
1
Draft,
do
not
cite
or
quote
Chapter
IX.
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