Health
Effects
Support
Document
for
Hexachlorobutadiene
Printed
on
Recycled
Paper
Health
Effects
Support
Document
for
Hexachlorobutadiene
U.
S.
Environmental
Protection
Agency
Office
of
Water
(
4304T)
Health
and
Ecological
Criteria
Division
Washington,
DC
20460
www.
epa.
gov/
safewater/
ccl/
pdf/
hexachlorobutadiene.
pdf
EPA
822­
R­
03­
002
February
2003
iii
HCBD
 
February
2003
FOREWORD
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1996,
requires
the
Administrator
of
the
Environmental
Protection
Agency
(
EPA)
to
establish
a
list
of
contaminants
to
aid
the
agency
in
regulatory
priority
setting
for
the
drinking
water
program.
In
addition,
SDWA
requires
EPA
to
make
regulatory
determinations
for
no
fewer
than
five
contaminants
by
August
2001.
The
criteria
used
to
determine
whether
or
not
to
regulate
a
chemical
on
the
CCL
are
the
following:

The
contaminant
may
have
an
adverse
effect
on
the
health
of
persons.

The
contaminant
is
known
to
occur
or
there
is
a
substantial
likelihood
that
the
contaminant
will
occur
in
public
water
systems
with
a
frequency
and
at
levels
of
public
health
concern.

In
the
sole
judgment
of
the
administrator,
regulation
of
such
contaminant
presents
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.

The
Agency's
findings
for
the
three
criteria
are
used
in
making
a
determination
to
regulate
a
contaminant.
The
Agency
may
determine
that
there
is
no
need
for
regulation
when
a
contaminant
fails
to
meet
one
of
the
criteria.

This
document
provides
the
health
effects
basis
for
the
preliminary
regulatory
determination
for
hexachlorobutadiene.
In
arriving
at
the
preliminary
regulatory
determination,
data
on
toxicokinetics,
human
exposure,
acute
and
chronic
toxicity
to
animals
and
humans,
epidemiology,
and
mechanisms
of
toxicity
were
evaluated.
In
order
to
avoid
duplication
of
effort,
information
from
the
following
risk
assessments
by
the
EPA
and
other
government
agencies
was
used
in
development
of
this
document.

U.
S.
EPA.
1991a.
Drinking
Water
Health
Advisory:
Hexachlorobutadiene.
In:
Volatile
Organic
Compounds.
United
States
Environmental
Protection
Agency,
Office
of
Drinking
Water.
Lewis
Publishers.
Ann
Arbor,
Michigan.

ATSDR.
1994.
Toxicological
Profile
for
Hexachlorobutadiene.
Agency
for
Toxic
Substances
and
Disease
Registry,
Department
of
Health
and
Human
Services.

U.
S.
EPA,
1998a.
Draft
Ambient
Water
Quality
Criteria
for
the
Protection
of
Human
Health.
Office
of
Water.
EPA
822­
R­
98­
004.

Information
from
the
published
risk
assessments
was
supplemented
with
information
from
recent
studies
of
hexachlorobutadiene
identified
by
literature
searches
conducted
in
1999
and
2000
and
the
primary
references
for
key
studies.

Generally
a
Reference
Dose
(
RfD)
is
provided
as
the
assessment
of
long­
term
toxic
effects
other
than
carcinogenicity.
RfD
determination
assumes
that
thresholds
exist
for
certain
toxic
effects
such
as
cellular
necrosis.
It
is
expressed
in
terms
of
milligrams
per
kilogram
per
day
(
mg/
kg­
day).
In
general,
the
RfD
is
an
estimate
(
with
uncertainty
spanning
perhaps
an
order
of
magnitude)
of
a
iv
HCBD
 
February
2003
daily
exposure
to
the
human
population
(
including
sensitive
subgroups)
that
is
likely
to
be
without
an
appreciable
risk
of
deleterious
effects
during
a
lifetime.

The
carcinogenicity
assessment
for
hexachlorobutadiene
includes
a
formal
hazard
identification
as
well
as
a
quantitative
dose­
response
assessment
of
the
risk
from
oral
exposure.
Hazard
identification
is
a
weight­
of­
evidence
judgment
of
the
likelihood
that
the
agent
is
a
human
carcinogen
via
the
oral
route
and
the
conditions
under
which
the
carcinogenic
effects
may
be
expressed.

Guidelines
that
were
used
in
the
development
of
this
assessment
may
include
the
following:
the
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1986a),
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
USEPA,
1999c),
Guidelines
for
the
Health
Risk
Assessment
of
Chemical
Mixtures
(
U.
S.
EPA,
1986b),
Guidelines
for
Mutagenicity
Risk
Assessment
(
U.
S.
EPA,
1986c),
Guidelines
for
Developmental
Toxicity
Risk
Assessment
(
U.
S.
EPA,
1991b),
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
1996a),
Guidelines
for
Reproductive
Toxicity
Risk
Assessment
(
U.
S.
EPA,
1996b),
and
Guidelines
for
Neurotoxicity
Risk
Assessment
(
U.
S.
EPA,
1998b);
Recommendations
for
and
Documentation
of
Biological
Values
for
Use
in
Risk
Assessment
(
U.
S.
EPA,
1988);
Use
of
the
Benchmark
Dose
Approach
in
Health
Risk
Assessment
(
U.
S.
EPA,
1995);
Science
Policy
Council
Handbook:
Peer
Review
(
U.
S.
EPA,
1998c);
and
Memorandum
from
EPA
Administrator,
Carol
Browner,
dated
March
21,
1995.

The
chapter
on
occurrence
and
exposure
to
hexachlorobutadiene
through
potable
water
was
developed
by
the
Office
of
Ground
Water
and
Drinking
Water.
It
is
based
primarily
on
unregulated
contaminant
monitoring
(
UCM)
data
collected
under
SDWA.
The
UCM
data
are
supplemented
with
ambient
water
data
as
well
as
information
on
production,
use,
and
discharge.
v
HCBD
 
February
2003
ACKNOWLEDGMENTS
This
document
was
prepared
under
the
U.
S.
EPA
Contract
No.
68­
C­
02­
009,
Work
Assignment
No.
B­
13
with
ICF
Consulting,
Fairfax,
VA.
The
Lead
U.
S.
EPA
Scientist
is
Diana
Wong,
Ph.
D.,
DABT,
Health
and
Ecological
Criteria
Division,
Office
of
Science
and
Technology,
Office
of
Water.
vi
HCBD
 
February
2003
TABLE
OF
CONTENTS
FOREWORD
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ACKNOWLEDGMENTS
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LIST
OF
TABLES
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ix
LIST
OF
FIGURES
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1.0
EXECUTIVE
SUMMARY
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1­
1
2.0
IDENTITY:
PHYSICAL
AND
CHEMICAL
PROPERTIES
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2­
1
3.0
USES
AND
ENVIRONMENTAL
FATE
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3­
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3.1
Uses
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3­
1
3.2
Release
to
the
Environment
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3­
3
3.3
Fate
in
Air
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3­
3
3.4
Fate
in
Water
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3­
3
3.5
Fate
in
Soil
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3­
4
4.0
EXPOSURE
FROM
DRINKING
WATER
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4­
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4.1
Ambient
Occurrence
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4­
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4.1.1
Data
Sources
and
Methods
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4­
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4.1.2
Results
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4­
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4.2
Drinking
Water
Occurrence
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4­
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4.2.1
Data
Sources,
Data
Quality,
and
Analytical
Methods
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4­
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4.2.2
Results
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4­
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4.3
Conclusions
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4­
17
5.0
EXPOSURE
FROM
MEDIA
OTHER
THAN
WATER
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5­
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5.1
Exposure
from
Food
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5­
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5.1.1
Concentrations
in
Non­
Fish
Food
Items
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5­
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5.1.2
Concentrations
in
Fish
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5­
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5.1.3
Intake
of
HCBD
from
Food
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5­
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5.2
Exposure
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Air
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5­
4
5.2.1
Concentration
of
HCBD
in
Air
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5­
4
5.2.2
Intake
of
HCBD
from
Air
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5­
5
5.3
Exposure
from
Soil
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5­
6
5.3.1
Concentration
of
HCBD
in
Soil
and
Sediment
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5­
6
5.3.2
Intake
of
HCBD
from
Soil
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5­
6
5.4
Other
Residential
Exposures
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5­
6
5.5
Summary
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5­
6
6.0
TOXICOKINETICS
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6­
1
vii
HCBD
 
February
2003
6.1
Absorption
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6­
1
6.2
Distribution
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6­
2
6.3
Metabolism
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6­
3
6.4
Excretion
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6­
7
7.0
HAZARD
IDENTIFICATION
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7­
1
7.1
Human
Effects
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7­
1
7.1.1
Short­
Term
Studies
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7­
1
7.1.2
Long­
Term
and
Epidemiological
Studies
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7­
1
7.2
Animal
Studies
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7­
2
7.2.1
Acute
Toxicity
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7­
2
7.2.2
Short­
Term
Studies
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7­
6
7.2.3
Subchronic
Studies
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7­
8
7.2.4
Neurotoxicity
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7­
10
7.2.5
Developmental/
Reproductive
Toxicity
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7­
11
7.2.6
Chronic
Toxicity
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7­
13
7.2.7
Carcinogenicity
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7­
14
7.3
Other
Key
Data
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7­
15
7.3.1
Mutagenicity/
Genotoxicity
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7­
15
7.3.2
Immunotoxicity
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7­
22
7.3.3
Hormonal
Disruption
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7­
23
7.3.4
Physiological
or
Mechanistic
Studies
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7­
23
7.3.5
Structure­
Activity
Relationship
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7­
27
7.4
Hazard
Characterization
.
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7­
28
7.4.1
Synthesis
and
Evaluation
of
Major
Noncancer
Effects
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7­
28
7.4.2
Synthesis
and
Evaluation
of
Carcinogenic
Effects
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7­
34
7.4.3
Mode
of
Action
and
Implications
in
Cancer
Assessment
.
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7­
34
7.4.4
Weight
of
Evidence
Evaluation
for
Carcinogenicity
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7­
36
7.4.5
Sensitive
Populations
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7­
37
8.0
DOSE­
RESPONSE
ASSESSMENT
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8­
1
8.1
Dose­
Response
for
Noncancer
Effects
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8­
1
8.1.1
RfD
Determination
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8­
1
8.1.2.
RfC
Determination
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8­
5
8.2
Dose­
Response
for
Cancer
Effects
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8­
5
8.2.1
Choice
of
Study
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8­
5
8.2.2
Dose­
Response
Characterization
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8­
6
8.2.3
Extrapolation
Model
and
Rationale
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8­
10
8.2.4
Cancer
Potency
and
Unit
Risk
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.
8­
11
8.2.5
Discussion
of
Confidence
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.
8­
12
9.0
REGULATORY
DETERMINATION
AND
CHARACTERIZATION
OF
RISK
FROM
DRINKING
WATER
.
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.
9­
1
9.1
Regulatory
Determination
for
Chemicals
on
the
CCL
.
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.
9­
1
9.1.1
Criteria
for
Regulatory
Determination
.
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.
.
9­
1
viii
HCBD
 
February
2003
9.1.2
National
Drinking
Water
Advisory
Council
Recommendations
.
.
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9­
2
9.2
Health
Effects
.
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.
9­
2
9.2.1
Health
Criterion
Conclusion
.
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.
9­
3
9.2.2
Hazard
Characterization
and
Mode
of
Action
Implications
.
.
.
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.
.
9­
3
9.2.3
Dose­
Response
Characterization
and
Implications
in
Risk
Assessment
.
.
.
.
.
.
9­
4
9.3
Occurrence
in
Public
Water
Systems
.
.
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.
9­
8
9.3.1
Occurrence
Criterion
Conclusion
.
.
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.
9­
8
9.3.2
Monitoring
Data
.
.
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.
9­
8
9.3.3
Use
and
Fate
Data
.
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.
9­
9
9.4
Risk
Reduction
.
.
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.
9­
10
9.4.1
Risk
Reduction
Criterion
Conclusion
.
.
.
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.
9­
11
9.4.2
Exposed
Population
Estimates
.
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.
9­
11
9.4.3
Relative
Source
Contribution
.
.
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9­
11
9.4.4
Sensitive
Populations
.
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9­
13
9.5
Regulatory
Determination
Summary
.
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9­
13
10.0
REFERENCES
.
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10­
1
APPENDIX
A:
Abbreviations
and
Acronyms
.
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A­
1
APPENDIX
B:
Round
1
and
Round
2
Occurrence
Data
Tables
for
Hexachlorobutadiene
.
.
.
B­
1
ix
HCBD
 
February
2003
LIST
OF
TABLES
Table
2­
1.
Chemical
and
Physical
Properties
of
Hexachlorobutadiene.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
2­
2
Table
3­
1.
Environmental
Releases
(
in
pounds)
for
Hexachlorobutadiene
in
the
United
States,
1988
 
1998.
.
.
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.
3­
2
Table
4­
1.
Cross­
section
States
for
Round
1
(
24
States)
and
Round
2
(
20
States).
.
.
.
.
.
.
.
.
.
4­
7
Table
4­
2.
Summary
Occurrence
Statistics
for
Hexachlorobutadiene.
.
.
.
.
.
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.
.
4­
12
Table
5­
1.
HCBD
Tissue
Concentration
in
Fish
Collected
Near
Four
Chemical
Manufacturing
Plants.
.
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.
5­
3
Table
5­
2.
Summary
of
Concentration
Data
and
Exposure
Estimates
for
Media
Other
Than
Water.
.
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.
5­
7
Table
7­
1.
Histopathological
Findings
in
Adult
Rats
Fed
Diets
containing
Hexachlorobutadiene.
.
.
.
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.
7­
9
Table
7­
2.
Mutagenicity
of
HCBD
in
Salmonella
typhimurium
Test
Systems.
.
.
.
.
.
.
.
.
.
.
.
7­
17
Table
7­
3.
Mutagenicity
of
HCBD
Metabolites.
.
.
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.
7­
19
Table
7­
4.
Genotoxicity
of
HCBD
in
Eukaryotic
Assay
Systems.
.
.
.
.
.
.
.
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.
7­
21
Table
7­
5.
Summary
of
Principal
HCBD
Toxicity
Studies.
.
.
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.
7­
32
Table
8­
1.
Incidence
of
Renal
Tubular
Regenerative
Response
in
Mice
Treated
with
HCBD
for
13
Weeks.
.
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.
8­
3
Table
8­
2.
Benchmark
Dose
Estimates
from
NTP
(
1991)
Female
Mouse
Renal
Tubular
Regeneration
Response.
.
.
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.
8­
4
Table
8­
3.
Incidence
of
Renal
Tubular
Neoplasms
in
Rats
Treated
with
HCBD
for
2
Years.
.
8­
6
Table
8­
4.
Dose­
Related
Changes
in
the
Rodent
Kidney
after
Oral
Exposure
to
HCBD,
Chronic
Study
­
Rat
(
Kociba
et
al.,
1977).
.
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.
8­
7
Table
8­
5.
Summary
of
Cancer
Risk
Values
for
HCBD.
.
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.
.
8­
13
Table
9­
1.
Dose­
Response
Information
from
Several
Key
Studies
of
HCBD
Toxicity
(
Oral
Exposure).
.
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.
9­
7
Table
9­
2.
National
Population
Estimates
for
HCBD
Exposure
via
Drinking
Water.
.
.
.
.
.
.
9­
12
Table
9­
3.
Comparison
of
Average
Daily
Intakes
from
Drinking
Water
and
Other
Media
.
.
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.
.
9­
13
Table
9­
4.
Ratios
a
of
Exposures
from
Various
Media
to
Exposures
from
Drinking
Water.
.
.
.
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.
9­
13
x
HCBD
 
February
2003
LIST
OF
FIGURES
Figure
2­
1.
Chemical
Structure
of
Hexachlorobutadiene.
.
.
.
.
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.
.
.
2­
1
Figure
4­
1.
Geographic
Distribution
of
Cross­
section
States
for
Round
1
(
left)
and
Round
2
(
right).
.
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.
.
4­
7
Figure
4­
2.
States
with
PWSs
with
Detections
of
Hexachlorobutadiene
for
all
States
with
Data
in
URCIS
(
Round
1)
and
SDWIS/
FED
(
Round
2).
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
4­
15
Figure
4­
3.
States
with
PWSs
with
Detections
of
Hexachlorobutadiene
(
any
PWSs
with
results
greater
than
the
Minimum
Reporting
Level
[
MRL])
for
Round
1
(
above)
and
Round
2
(
below)
Cross­
section
States.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
.
4­
15
Figure
4­
4.
Cross­
section
States
(
Round
1
and
Round
2
combined)
with
PWSs
with
Detections
of
Hexachlorobutadiene
(
above)
and
concentrations
greater
than
the
Health
Reference
Level
(
HRL;
below).
.
.
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.
.
4­
17
Figure
6­
1.
Proposed
Pathways
for
Hexachlorobutadiene
Metabolism.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
6­
4
Figure
8­
1.
Benchmark
Dose
Estimate
Using
Weibull
Model.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
8­
2
Figure
8­
2.
Renal
Tumor
Dose
Response
Curves
.
.
.
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.
.
8­
10
HCBD
 
February
2003
1­
1
1.0
EXECUTIVE
SUMMARY
The
U.
S.
Environmental
Protection
Agency
(
EPA)
has
prepared
this
Health
Effects
Support
Document
for
Hexachlorobutadiene
(
HCBD)
to
assist
in
determining
whether
to
regulate
HCBD
with
a
National
Primary
Drinking
Water
Regulation
(
NPDWR).
The
available
data
on
occurrence,
exposure,
and
other
risk
considerations
suggest
that,
because
HCBD
does
not
occur
in
public
water
systems
at
frequencies
and
levels
of
public
health
concern,
regulating
HCBD
will
not
present
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.
EPA
presents
its
determination
and
data
analysis
in
the
Federal
Register
Notice
covering
the
Contaminant
Candidate
List
(
CCL)
regulatory
determinations.

HCBD
(
Chemical
Abstracts
Services
Registry
Number
87­
68­
3)
is
a
colorless
liquid
at
room
temperature.
It
is
poorly
soluble
in
water,
and
has
a
high
affinity
for
organic
particulate
matter.
HCBD
has
never
been
specifically
manufactured
as
a
commercial
product
in
the
United
States.
However,
significant
quantities
of
hexachlorobutadiene
are
generated
in
the
United
States
as
waste
by­
product
from
the
chlorination
of
hydrocarbons.
The
chemical
is
used
as
an
intermediate
product
in
rubber
manufacturing
and
chlorofluorocarbon
and
lubricant
production,
as
well
as
for
transformer
and
hydraulic
fluids,
fluid
for
gyroscopes,
heat
transfer
liquid,
solvents,
laboratory
reagents,
and
as
a
wash
liquor
for
removing
C
4
and
higher
hydrocarbons.
Hexachlorobutadiene
has
also
been
used
as
a
fumigant
in
some
overseas
countries.
Some
of
the
chemical
properties
for
hexachlorobutadiene
(
CAS#
87­
68­
3)
include
the
following:
water
solubility
=
2­
2.55
mg/
L;
vapor
pressure
(
25
°
C)
=
0.15
mmHg;
Log
K
ow
=
4.78;
and
Log
K
oc
=
3.67.

Emissions
into
air
is
the
major
pathway
of
release.
For
hexachlorobutadiene,
air
emissions
constitute
most
of
the
on­
site
releases.
Hexachlorobutadiene
is
listed
as
a
toxic
release
inventory
(
TRI)
chemical.
It
is
included
in
the
Agency
for
Toxic
Substances
and
Disease
Registry's
(
ATSDR)
Hazardous
Substance
Release
and
Health
Effects
Database
(
HazDat)
and
has
been
detected
in
site
samples
in
fifteen
States:
AL,
AZ,
CA,
CT,
IA,
LA,
MI,
MN,
NJ,
NY,
OH,
PA,
RI,
SC,
WA
(
ATSDR,
2000).

Ambient
air
concentration
data
are
available
from
Shah
and
Heyerdahl
(
1988).
The
mean
and
median
of
all
ambient
concentrations
were
0.42
µ
g/
m3
and
0.04
µ
g/
m3,
respectively.
Air
intake
for
adults
is
estimated
to
be
1.2
×
10­
4
mg/
kg­
day
using
the
mean
air
concentration,
and
inhalation
is
the
main
pathway
of
exposure
(
U.
S.
EPA,
1998a).
Hexachlorobutadiene
is
not
found
in
non­
fish
dietary
foods
for
the
majority
of
regions
in
the
US.
It
was
detected
in
fish
at
3%
of
362
sites
sampled.
The
mean
fish
concentration
at
all
sites
was
0.6
ng/
g
(
Kuehl
et
al.,
1994).
An
estimate
of
adult
exposure
via
fish
consumption
is
1.54
×
10­
7
mg/
kg­
day
(
U.
S.
EPA,
1998a).

Cross­
sectional
monitoring
data
from
two
rounds
of
sampling
conducted
under
EPA's
Unregulated
Contaminant
Monitoring
(
UCM)
program
indicate
that
the
frequency
of
detection
of
HCBD
in
public
water
systems
(
PWSs)
is
low.
Round
1,
conducted
from
1987
to
1992
in
24
States,
detected
HCBD
at
levels
above
the
minimum
reporting
level
(
MRL)
of
0.5
µ
g/
L
for
0.35%
of
the
PWSs,
while
Round
2,
conducted
from
1993
to
1997
in
20
States,
detected
HCBD
at
levels
above
the
MRL
for
0.18%
of
the
PWSs.
The
health
reference
level
(
HRL)
of
0.9
:
g/
L
for
HCBD
is
a
preliminary
health
effect
level
used
in
the
UCM
analysis
(
U.
S.
EPA,
2001c).
It
is
the
concentration
HCBD
 
February
2003
1­
2
corresponding
to
10­
6
incremental
cancer
risk,
calculated
from
the
slope
factor
using
the
linear
method.
In
that
analysis,
0.114%
of
the
PWSs
in
Round
1
(
74
systems)
exceeded
the
HRL,
while
0.018%
(
11
systems)
in
Round
2
exceeded
the
HRL.
The
United
States
Geological
Survey's
National
Ambient
Water
Quality
Assessment
(
NAWQA)
program
did
not
detect
HCBD
in
the
ground
water
or
well
water
samples
surveyed.
When
average
daily
drinking
water
intakes
for
HCBD
are
compared
with
intakes
from
food,
air
and
soil,
drinking
water
accounts
for
a
relatively
small
proportion
of
total
HCBD
intake.
This
drinking
water
intake,
however,
does
not
include
other
uses
for
potable
water
such
as
bathing
and
showering.

Hexachlorobutadiene
(
HCBD)
is
readily
absorbed
following
oral
administration
in
rats.
Following
gastrointestinal
uptake,
HCBD
and
its
metabolites
distribute
preferentially
to
the
kidney,
liver,
adipose
tissue
and
the
brain
(
Reichert
et
al.,
1985).
The
primary
pathway
for
HCBD
metabolism
is
conjugation
with
glutathione,
with
subsequent
conversion
to
a
cysteine
conjugate.
Activation
of
the
cysteine
conjugate
by
$­
lyase
yields
a
highly
reactive
thioketene
intermediate.
Covalent
binding
of
this
thioketene
to
DNA,
proteins
and
other
macromolecules
is
considered
to
be
the
mechanism
responsible
for
the
observed
cytotoxic
and
mutagenic
effects
of
HCBD
and
its
metabolites.
Evidence
exists
for
a
male
specific
metabolic
pathway
in
rats
(
Birner
et
al.,
1995).
The
primary
routes
for
elimination
of
absorbed
HCBD
are
urinary
and
fecal
excretion;
a
small
amount
of
absorbed
HCBD
is
oxidized
to
carbon
dioxide
(
U.
S.
EPA,
1991;
U.
S.
EPA,
1999;
ATSDR,
1994).

There
are
no
reliable
dose­
response
data
for
humans
exposed
to
HCBD.
There
is
no
information
available
to
determine
the
carcinogenic
potential
of
HCBD
exposure
in
humans.
Studies
in
animals
show
a
selective
adverse
effect
of
HCBD
on
the
kidney,
specifically
the
proximal
tubule.
Subchronic
(
NTP,
1991)
and
chronic
(
Kociba
et
al.,
1977)
studies
in
rodents
present
a
clear
picture
of
dose­
related
renal
damage
at
2
mg/
kg­
day
and
above,
with
possible
effects
at
0.2
mg/
kg­
day.
Progressive
events
over
time
include
changes
in
kidney
weight,
renal
tubular
degeneration
and
regeneration,
hyperplasia,
focal
adenomatous
proliferation,
and
renal
tumor
formation.
One
subacute
inhalation
study
also
found
enlarged
adrenals
and
degeneration
of
adrenal
cortex
at
250
ppm
(
Gage,
1970).
Developmental
effects
were
also
associated
with
hexachlorobutadiene
exposure
in
animals
(
Harleman
and
Seinen,
1979).
However,
these
effects
were
observed
at
higher
doses
than
required
to
produce
renal
toxicity.
Pups
with
lower
birth
weights
and
reduced
body
weight
gain
were
reported
at
maternal
dose
of
8.1­
15
mg/
kg­
day
in
rats
(
Badaeva,
1983;
Harleman
and
Seinen,
1979).
In
the
presence
of
metabolic
activation,
HCBD
and
its
reactive
metabolites
are
mutagenic
in
some
(
Simmon,
1977;
Reichert
et
al.,
1984;
Reichert
and
Schutz,
1986;
Wild
et
al.,
1986)
but
not
all
studies.
Only
one
study
of
lifetime
oral
exposure
to
hexachlorobutadiene
was
located
(
Kociba
et
al.,
1977).
At
20
mg/
kg­
day,
benign
and
malignant
renal
tumors
were
seen
in
approximately
23%
(
9/
39)
of
the
male
rats,
and
15%
(
6/
40)
of
the
female
rats.
This
dose
exceeded
the
maximum
tolerated
dose
at
which
increased
mortality,
severe
renal
toxicity,
and
significant
weight
loss
were
also
observed.
There
were
no
tumors
in
the
second
highest
dose
of
2
mg/
kg­
day
in
this
study.
The
conclusion
from
the
dose
response
analysis
is
that
hexachlorobutadiene
is
carcinogenic
only
in
the
kidney
following
ingestion
at
cytotoxic
doses
in
the
rat.

The
nephrotoxicity
of
HCBD
in
rodents
is
dependent
on
a
multi­
step
bioactivation
mechanism
involving
both
kidney
and
liver
enzymes.
The
ultimate
step
in
this
pathway
is
a
$­
lyase
mediated
degradation
of
a
HCBD
metabolite
that
generates
a
highly
reactive
thioketene
in
proximal
tubule
cells.
In
vitro
studies
suggest
that
cortical
mitochondria
are
the
critical
subcellular
target
for
HCBD
 
February
2003
1­
3
toxicity.
Covalent
binding
of
this
reactive
thioketene
to
cellular
macromolecules
(
e.
g.
proteins,
mitochondrial
DNA)
and
the
resultant
mitochondrial
dysfunction
is
believed
to
underlie
the
development
of
renal
cytotoxicity
and
renal
tubular
adenomas
and
carcinomas
in
rodents.

Limited
in
vitro
studies
suggest
humans
have
the
ability
to
metabolize
HCBD.
The
activity
of
HCBD
metabolizing
enzymes,
particularly
renal
$­
lyase,
may
be
many
fold
lower
in
humans
than
the
corresponding
enzymes
in
rats
(
Lock,
1994;
Lash
et
al.,
1990;
Anders
and
Dekant,
1998);
although
the
activity
of
glutathione
S­
transferase
was
not
different
between
rodents
and
humans
(
Dekant
et
al.,
1998).
Due
to
the
fact
that
lower
levels
of
reactive
metabolites
would
be
assumed
to
form,
there
would
be
less
concern
for
toxicity
in
humans.

The
primary
target
organ
for
HCBD
is
the
kidney.
Individuals
with
preexisting
kidney
damage
may
represent
a
potentially
sensitive
subpopulation
for
hexachlorobutadiene
health
effects.
Studies
in
animals
showed
that
the
young
rats
and
mice
were
more
sensitive
to
the
acute
effects
of
oral
HCBD
than
adults
(
Hook
et
al.,
1983;
Lock
et
al.,
1984).
Those
data
may
suggest
that
infants
may
potentially
be
more
susceptible
to
hexachlorobutadiene
toxicity.

Three
key
studies
which
have
been
used
in
determining
points
of
departure
for
the
reference
dose
(
RfD)
are
the
Kociba
et
al.
(
1977)
study
and
Schwetz
et
al.
(
1977)
studies
in
rats
and
the
NTP
(
1991)
study
in
mice.
In
the
Kociba
(
1977)
study,
male
and
female
Sprague­
Dawley
rats
were
fed
diets
that
contained
0,
0.2,
2,
or
20
mg/
kg­
day
HCBD
for
22
months
(
males)
or
24
months
(
females).
Non­
neoplastic
effects,
including
increased
coproporphyrin
excretion
and
microscopic
renal
tubule
epithelial
hyperplasia,
were
seen
at
the
two
high
doses,
while
renal
tubule
adenomas
and
carcinomas
were
found
only
at
the
highest
dose.
In
the
Schwetz
et
al.
(
1977)
study,
male
and
female
Sprague­
Dawley
rats
were
fed
a
diet
containing
0.2,
2.0,
or
20
mg/
kg­
day
HCBD
for
evaluation
of
reproductive
effects.
HCBD
was
provided
in
the
diet
before
and
during
mating,
and
throughout
gestation
and
lactation,
for
a
total
study
duration
of
148
days.
At
necropsy,
relative
kidney
weights
were
increased
in
high­
dose
males
and
females.
Relative
liver
weight
was
increased
in
high­
dose
males,
and
relative
brain
weight
was
increased
in
high­
dose
females.
The
kidneys
of
males
at
the
two
high
doses
were
roughened
and
had
a
mottled
cortex.
Histopathological
examination
revealed
dose­
related
increases
in
tubular
dilation
and
regeneration
in
animals
at
the
two
high
doses.
In
the
13­
week
NTP
(
1991)
study,
B6C3F
1
mice
were
fed
a
diet
containing
0,
0.1,
0.4,
1.5,
4.9
or
16.8
mg/
kg­
day
for
males
and
0,
0.2,
0.5,
1.8,
4.5
or
19.2
mg/
kg­
day
for
females.
Reduced
body
weight
gain
was
reported
in
males
at
the
two
high
doses
and
in
females
at
the
highest
dose.
Relative
kidney
weight
was
decreased
in
the
three
high
dose
groups
for
males,
and
in
females
in
the
highest
dose
group.
High­
dose
males
also
exhibited
decreased
relative
heart
weight.
Necropsy
revealed
treatmentrelated
increases
in
renal
tubular
cell
regeneration
in
all
dose
groups
in
female
mice
and
in
the
two
high
dose
groups
in
male
mice.

A
previous
RfD
for
hexachlorobutadiene
was
2
x
10­
4
mg/
kg­
day
(
EPA,
1998a).
It
was
derived
from
a
NOAEL
of
0.2
mg/
kg­
day
for
renal
tubular
epithelial
cell
degeneration
and
regeneration
from
the
Kociba
et
al.
(
1977)
study
on
rats
and
supported
by
the
NTP
(
1991)
study
on
mice.
A
composite
uncertainty
factor
(
UF)
of
1,000
was
used
in
the
derivation
of
the
RfD.

The
RfD
for
hexachlorobutadiene
is
3
x
10­
4
mg/
kg­
day.
This
is
derived
from
the
BMDL
of
0.1
mg/
kg­
day
calculated
from
a
BMD
analysis
of
renal
tubular
epithelial
cell
regeneration
from
the
HCBD
 
February
2003
1­
4
NTP
(
1991)
study,
with
support
from
the
Kociba
(
1977)
and
Schwetz
(
1977)
studies
in
rats.
A
composite
UF
of
300
is
now
used
in
the
derivation
of
the
RfD.
The
composite
uncertainty
factor
includes
a
factor
of
10
to
account
for
extrapolation
from
animals
to
humans;
a
factor
of
10
for
protection
of
sensitive
subpopulation;
and
a
factor
of
3
for
database
deficiencies
(
lack
of
a
2­
generation
reproductive
study
and
developmental
toxicity
studies
in
only
one
species).
In
accordance
with
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1986),
HCBD
is
classified
as
a
Group
C
(
possible
human)
carcinogen.
Under
EPA's
1999
draft
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999c),
HCBD
is
classified
as
likely
to
be
carcinogenic
to
humans
by
the
oral
route
of
exposure.
Its
carcinogenic
potential
by
the
inhalation
and
dermal
routes
of
exposure
is
classified
as
cannot
be
determined
because
there
are
inadequate
data
to
perform
an
assessment.
Two
different
approaches
were
used
for
dose­
response
extrapolation
to
estimate
excess
human
cancer
risk
for
HCBD
exposure
from
the
rodent
data.
The
default
linearized
multi­
stage
model
calculated
a
slope
factor
of
4
×
10­
2
(
mg/
kg­
day)­
1,
and
a
unit
risk
of
1.1
×
10­
6
per
:
g/
L.
For
the
nonlinear
approach,
the
RfD
of
3
×
10­
4
mg/
kg­
day
is
used
for
the
protection
of
cancer
effect.
In
consideration
of
the
overall
evidence,
including
the
possible
genotoxicity
of
HCBD
metabolites,
the
linear
approach
is
recommended
by
EPA.
HCBD
 
February
2003
2­
1
Hexachlorobutadiene
2.0
IDENTITY:
PHYSICAL
AND
CHEMICAL
PROPERTIES
The
chemical
and
physical
properties
of
hexachlorobutadiene
(
HCBD)
are
summarized
in
Table
2­
1.
Synonyms
for
this
chemical
include
perchlorobutadiene;
1,1,2,3,4,4­
hexachloro­
1,3­
butadiene;
1,3­
hexachlorobutadiene;
Dolen­
Pur;
and
GP­
40­
66:
120.

HCBD
is
a
colorless
liquid
at
room
temperature
with
a
mild
turpentine­
like
odor
(
HSDB,
2000).
HCBD
is
poorly
soluble
in
water,
but
is
miscible
in
ethanol
and
ether
(
HSDB,
2000).
HCBD
has
a
relatively
low
vapor
pressure
of
0.15
mm
Hg
at
25
°
C
(
U.
S.
EPA,
1991a).
An
odor
threshold
of
0.006
mg/
L
has
been
reported
for
HCBD
in
water
(
U.
S.
EPA,
1980),
and
an
air
odor
threshold
of
12
mg/
m3
(
Ruth,
1986).
The
Occupational
Safety
and
Health
Administration
(
OSHA)
Permissible
Exposure
Limit
(
PEL)
is
0.21
mg/
m3
(
OSHA,
1989),
making
odor
a
poor
warning
characteristic
for
HCBD.
HCBD
is
characterized
by
high
log
K
oc
and
log
K
ow
values,
3.67
and
4.78,
respectively
(
ATSDR,
1994),
reflecting
properties
which
strongly
influence
its
behavior
and
fate
in
environmental
media.
The
chemical
structure
of
HCBD
is
shown
in
Figure
2­
1.

Figure
2­
1.
Chemical
Structure
of
Hexachlorobutadiene.
HCBD
 
February
2003
2­
2
Table
2­
1.
Chemical
and
Physical
Properties
of
Hexachlorobutadiene.

Property
Information
Chemical
Formula
C
4
Cl
6
Molecular
Weight
260.76
Synonyms
HCBD;
Perchlorobutadiene;
Hexachlorbutadiene;
1,1,2,3,4,4­
Hexachloro­
1,3­
butadiene;
1,3­
Hexachlorobutadiene;
Dolen­
Pur;
GP­
40­
66:
120NIOSH
Registry
of
Toxic
Effects
of
Chemical
Substances
(
RTECS)
No.
EJ0700000U.
S.
EPA
Hazardous
Waste
No.
U128Oil
and
Hazardous
Materials/
Technical
Assistance
Data
System
(
OHM/
TADS)
No.
OHM
8100011Hazardous
Substances
Data
Bank
(
HSDB)
No.
2870
Boiling
Point
(
at
760
mm
Hg)
215oC
Melting
Point
­
21oC
Vapor
Pressure
(
at
25oC)
0.15
mm
Hg
Density
(
at
20oC)
1.55
g/
cm3
Water
Solubility
(
at
20oC)
2
 
2.55
mg/
L
Organic
Solvents
Ethanol,
Ether
Partition
Coefficients
Log
K
ow
4.78
Log
K
oc
3.67
Odor
Threshold
(
air)
Odor
Threshold
(
water)
12.00
mg/
m3
0.006
mg/
L
Conversion
Factor
1
ppm
=
10.66
mg/
m3
1
mg/
m3
=
0.0938
ppm
Sources:
U.
S.
EPA
(
1980,
1991a);
ChemIDplus
(
2000);
HSDB
(
2000)
HCBD
 
February
2003
3­
1
3.0
USES
AND
ENVIRONMENTAL
FATE
3.1
Uses
HCBD
has
never
been
specifically
manufactured
as
a
commercial
product
in
the
United
States.
However,
significant
quantities
of
the
chemical
are
generated
in
the
U.
S.
as
waste
by­
product
from
the
chlorination
of
hydrocarbons
such
as
tri­
and
tetrachloroethylene
and
carbon
tetrachloride.
Lesser
quantities
have
been
imported
in
past
decades
for
commercial
purposes,
mostly
from
Germany.
Until
1975
the
most
important
commercial
use
of
HCBD
in
the
United
States
was
for
the
recovery
of
chlorine­
containing
gases
in
chlorine
plants.
Since
then,
HCBD
has
been
used
primarily
as
a
chemical
intermediate
in
the
production
of
rubber,
and
also
as
a
hydraulic
fluid,
a
fluid
for
gyroscopes,
a
heat
transfer
liquid,
a
solvent,
a
laboratory
reagent,
a
wash
liquor
for
removing
C
4
and
higher
hydrocarbons,
and
as
a
chemical
intermediate
in
the
production
of
chlorofluorocarbons
and
lubricants
(
ATSDR,
1995;
Howard,
1989).
The
chemical
has
also
been
used
as
a
fumigant
in
Russia,
France,
Italy,
Greece,
Spain,
and
Argentina,
although
use
in
the
European
Community
is
reported
to
have
ceased
(
van
de
Plassche
and
Schwegler,
2002).

Eight
million
pounds
of
HCBD
were
generated
as
a
waste
by­
product
in
the
U.
S.
in
1975,
with
0.1
million
pounds
released
into
the
environment.
By
1982,
the
annual
U.
S.
by­
product
generation
of
the
chemical
had
increased
to
28
million
pounds.
In
contrast,
the
annual
import
rate
of
HCBD
dropped
from
500,000
lbs/
yr
imported
annually
in
the
late
1970s,
to
145,000
lbs/
yr
imported
in
1981
(
ATSDR,
1994;
Howard,
1989).
Van
de
Plassche
and
Schwegler
(
2002)
report
that
all
commercial
production
in
Europe
has
ceased,
and
estimate
that
worldwide
commercial
production
has
dropped
from
10,000
tonnes
in
1982
to
virtually
nil
today.

3.2
Release
to
the
Environment
HCBD
is
listed
as
a
toxic
release
inventory
(
TRI)
chemical.
In
1986,
the
Emergency
Planning
and
Community
Right­
to­
Know
Act
(
EPCRA)
established
the
Toxic
Release
Inventory
(
TRI)
of
hazardous
chemicals.
Created
under
the
Superfund
Amendments
and
Reauthorization
Act
(
SARA)
of
1986,
EPCRA
is
also
sometimes
known
as
SARA
Title
III.
The
EPCRA
mandates
that
larger
facilities
publicly
report
when
TRI
chemicals
are
released
into
the
environment.
This
public
reporting
is
required
for
facilities
with
more
than
10
full­
time
employees
that
annually
manufacture
or
produce
more
than
25,000
pounds,
or
use
more
than
10,000
pounds,
of
TRI
chemical
(
U.
S.
EPA,
1996c,
2000a).

Under
these
conditions,
facilities
are
required
to
report
the
pounds
per
year
of
HCBD
released
into
the
environment
both
on­
and
off­
site.
The
on­
site
quantity
is
subdivided
into
air
emissions,
surface
water
discharges,
underground
injections,
and
releases
to
land
(
see
Table
3­
1).
For
HCBD,
air
emissions
constitute
most
of
the
on­
site
releases.
Also,
over
the
period
for
which
data
are
available
(
1988
 
1998),
surface
water
discharges
generally
increased,
peaked
in
1992
 
1993,
and
then
decreased
significantly
through
the
late
1990s.
The
TRI
data
for
HCBD
were
reported
from
eight
States
(
CA,
IL,
KS,
LA,
NJ,
NY,
TX,
UT);
however,
HCBD
contamination
has
often
been
found
in
remote
areas
far
from
apparent
physical
discharge
sources
(
U.
S.
EPA,
2000b;
Howard,
1989).
HCBD
 
February
2003
3­
2
Table
3­
1.
Environmental
Releases
(
in
pounds)
for
Hexachlorobutadiene
in
the
United
States,
1988
 
1998.

Year
On­
Site
Releases
Off­
Site
Releases
Total
On­
&
Off­
site
Releases
Air
Emissions
Surface
Water
Discharges
Underground
Injection
Releases
to
Land
1998
2,421
5
0
0
510
2,936
1997
1,415
9
299
0
200
1,923
1996
2,381
256
952
0
310
3,899
1995
3,310
661
434
0
252
4,657
1994
1,410
351
201
0
430
2,392
1993
1,747
1,200
520
0
12
3,479
1992
4,134
1,911
738
0
5
6,788
1991
3,410
681
200
2
4,263
8,556
1990
4,906
715
330
0
45
5,996
1989
4,628
622
330
1
26,343
31,924
1988
2,508
153
220
0
19,640
22,521
source:
U.
S.
EPA
(
2000b)

Although
the
TRI
data
can
be
useful
in
giving
a
general
idea
of
release
trends,
it
is
far
from
exhaustive
and
has
significant
limitations.
For
example,
only
industries
which
meet
TRI
criteria
(
at
least
10
full­
time
employees
and
manufacture
and
processing
of
quantities
exceeding
25,000
lbs/
yr,
or
use
of
more
than
10,000
lbs/
yr)
are
required
to
report
releases.
These
reporting
criteria
do
not
account
for
releases
from
smaller
industries.
Threshold
manufacture
and
processing
quantities
also
changed
from
1988
 
1990
(
dropping
from
75,000
lbs/
yr
in
1988
to
50,000
lbs/
yr
in
1989
to
its
current
25,000
lbs/
yr
in
1990)
creating
possibly
misleading
data
trends.
Finally,
the
TRI
data
is
meant
to
reflect
releases
and
should
not
be
used
to
estimate
general
exposure
to
a
chemical
(
U.
S.
EPA,
2000c,
d).

While
TRI
releases
were
reported
in
only
eight
States,
the
use
of
HCBD
is
widespread.
It
is
included
in
the
Agency
for
Toxic
Substances
and
Disease
Registry's
(
ATSDR)
Hazardous
Substance
Release
and
Health
Effects
Database
(
HazDat)
and
has
been
detected
in
site
samples
in
fourteen
States
(
AL,
AZ,
CT,
IA,
LA,
MI,
MN,
NJ,
NY,
OH,
PA,
RI,
SC,
WA;
ATSDR,
2000).
These
States
are
distributed
nationwide
and
include
11
States
and
two
regions
(
New
England
and
the
Pacific
Northwest)
not
reporting
TRI
releases
yet
manifesting
HCBD
detections
in
the
environment.

The
National
Priorities
List
(
NPL)
of
hazardous
waste
sites,
created
in
1980
by
the
Comprehensive
Environmental
Response,
Compensation
and
Liability
Act
(
CERCLA),
is
a
listing
of
some
of
the
most
health­
threatening
waste
sites
in
the
United
States.
HCBD
was
detected
in
eleven
NPL
sites
in
1999.
These
sites
are
located
in
eight
States:
AK,
CO,
IN,
LA,
NJ,
OH,
PA,
WA.
Again,
note
that
there
is
little
overlap
between
these
States
and
the
eight
TRI
reporting
States
(
U.
S.
EPA,
1999a).

In
summary,
although
HCBD
is
not
manufactured
in
the
United
States,
both
its
use
in
industry
and
occurrence
in
the
environment
are
widespread.
Significant
quantities
of
HCBD
are
generated
in
the
United
States
as
a
waste
by­
product,
and
smaller
quantities
are
imported
for
industrial
needs.
HCBD
is
present
in
hazardous
waste
sites
in
at
least
8
States
(
at
NPL
sites),
has
HCBD
 
February
2003
3­
3
been
detected
in
site
samples
in
at
least
14
States
(
listed
in
ATSDR's
HazDat),
and
has
been
released
into
the
environment
directly
in
at
least
8
States
(
based
on
TRI
data).

3.3
Fate
in
Air
HCBD
is
released
to
air
via
chemical
manufacturing
and
processing
and
by
waste
incineration
(
HSDB,
2000).
Modeling
and
monitoring
data
suggest
that
the
atmospheric
burden
of
HCBD
in
the
northern
hemisphere
is
approximately
3.2
million
kg/
yr
(
Class
and
Ballschmiter,
1987).
Dispersion
of
HCBD
in
the
atmosphere
has
been
confirmed
by
detection
of
HCBD
at
locations
distant
from
sources
of
release
(
WHO,
1994).
The
high
log
organic
carbon
partition
coefficient
(
log
K
oc)
of
HCBD
indicates
that
it
will
readily
adsorb
to
airborne
particulate
matter
with
a
high
organic
content.
Thus,
HCBD
in
air
is
found
both
as
a
vapor
and
in
association
with
atmospheric
particulates.

No
specific
information
is
available
on
the
transformation
and
degradation
of
HCBD
in
air.

3.4
Fate
in
Water
HCBD
is
released
to
surface
and
ground
water
via
industrial
effluents,
by
leaching
from
landfills
or
soil,
or
by
urban
runoff
(
ATSDR,
1994).
Sorption
to
sediments
and
suspended
particulates
is
an
important
factor
in
the
fate
of
HCBD
in
water
(
U.
S.
EPA,
1991a).
As
a
result
of
this
affinity
for
particulates
and
sediments,
HCBD­
contaminated
areas
will
usually
have
higher
sediment
concentrations
than
water
concentrations
of
the
chemical.
U.
S.
EPA
(
1976)
found
that
HCBD
concentrations
in
the
Mississippi
delta
water
were
<
2
:
g/
L,
while
concentrations
in
mud
or
soil
were
>
200
:
g/
L.
Leeuwangh
et
al.
(
1975)
observed
that
equilibration
of
initially
uncontaminated
sediment
with
HCBD­
contaminated
water
resulted
in
sediment
concentrations
100­
fold
greater
than
those
observed
in
the
water.

Volatilization
of
HCBD
from
water
to
air
also
occurs,
although
the
low
vapor
pressure
of
HCBD
(
0.15
mmHg
at
25
°
C)
suggests
that
this
process
may
occur
relatively
slowly
(
U.
S.
EPA,
1991a).
Limited
data
are
available
on
the
transformation
and
degradation
of
HCBD
in
water.
Under
aerobic
conditions
in
batch
culture,
complete
biodegradation
has
been
observed
to
occur
in
sewageinoculated
waters
after
seven
days
(
Tabak
et
al.,
1981).
These
data
suggest
that
HCBD
may
biodegrade
in
natural
waters.
In
contrast,
no
degradation
was
observed
under
anaerobic
conditions
in
a
separate
study
(
Johnson
and
Young,
1983).
No
data
were
available
on
hydrolysis
or
photolysis
of
HCBD
in
water.
Estimates
of
HCBD
half­
life
range
from
3
to
30
days
in
rivers
and
30
to
300
days
in
lakes
and
groundwater
(
Zoeteman
et
al.,
1980).

The
high
octanol­
water
partition
coefficient
(
K
ow)
of
HCBD
suggests
that
this
chemical
can
readily
partition
from
water
into
biota.
Laboratory
and
field
investigations
confirm
that
HCBD
has
bioaccumulation
potential
(
WHO,
1994).
Field­
measured
bioaccumulation
factors
range
from
46
to
27,780
(
U.
S.
EPA,
1999b).
No
evidence
for
biomagnification
has
been
observed
in
laboratory
or
field
studies
(
WHO,
1994).
HCBD
 
February
2003
3­
4
3.5
Fate
in
Soil
HCBD
can
be
released
to
soil
by
disposal
of
industrial
waste
in
landfill
operations
(
ATSDR,
1994).
Volatilization
from
soil
surfaces
is
expected
to
be
a
primary
process
for
loss
of
HCBD
from
soil
(
Tabak
et
al.,
1981).
However,
as
HCBD
readily
adsorbs
to
soil
organic
particles,
volatilization
from
highly
organic
soils
is
predicted
to
be
low
(
HSDB,
2000).

No
data
regarding
transformation
or
degradation
of
HCBD
in
soil
were
located.
Data
from
experiments
conducted
in
water
(
Tabak
et
al.,
1981)
suggest
that
biodegradation
will
occur
if
aerobic
conditions
are
present
(
HSDB,
2000).
Results
obtained
in
sludge
incubated
under
anaerobic
conditions
(
Johnson
and
Young,
1983)
suggest
that
biodegradation
will
not
occur
under
anaerobic
soil
conditions.
Soil
organic
matter
content
is
likely
to
be
an
important
factor
in
biodegradation
time,
since
adsorption
of
HCBD
to
organic
matter
will
significantly
decrease
its
bioavailability
to
microorganisms.
In
the
absence
of
significant
biodegradation
or
other
loss
processes,
persistence
of
HCBD
in
soil
may
allow
migration
of
the
compound
into
groundwater,
particularly
in
sandy
soils
(
U.
S.
EPA,
1984).
HCBD
 
February
2003
4­
1
4.0
EXPOSURE
FROM
DRINKING
WATER
This
section
of
the
report
examines
the
occurrence
of
HCBD
in
drinking
water.
While
no
complete
national
database
exists
of
unregulated
or
regulated
contaminants
in
drinking
water
from
public
water
systems
(
PWSs)
collected
under
SDWA,
this
report
aggregates
and
analyzes
existing
State
data
that
have
been
screened
for
quality,
completeness,
and
representativeness.
Populations
served
by
PWSs
exposed
to
HCBD
are
estimated,
and
the
occurrence
data
are
examined
for
regional
or
other
special
trends.
To
augment
the
incomplete
national
drinking
water
data
and
aid
in
the
evaluation
of
occurrence,
information
on
ambient
occurrence
of
HCBD
is
also
reviewed.

4.1
Ambient
Occurrence
To
understand
the
presence
of
a
chemical
in
the
environment,
an
examination
of
ambient
occurrence
is
useful.
In
a
drinking
water
context,
ambient
water
is
source
water
existing
in
surface
waters
and
aquifers
before
treatment.
The
most
comprehensive
and
nationally
representative
data
describing
ambient
water
quality
in
the
United
States
are
being
produced
through
the
United
States
Geological
Survey's
(
USGS)
National
Water
Quality
Assessment
(
NAWQA)
program.
NAWQA,
however,
is
a
relatively
young
program
and
complete
national
data
are
not
yet
available
from
their
entire
array
of
sites
across
the
nation.

4.1.1
Data
Sources
and
Methods
To
examine
water
quality
status
and
trends
in
the
United
States,
the
USGS
instituted
the
NAWQA
program
in
1991.
NAWQA
is
designed
and
implemented
in
such
a
manner
to
allow
consistency
and
comparison
between
representative
study
basins
located
around
the
country,
facilitating
interpretation
of
natural
and
anthropogenic
factors
affecting
water
quality
(
Leahy
and
Thompson,
1994).

The
NAWQA
program
consists
of
59
significant
watersheds
and
aquifers
referred
to
as
"
study
units."
The
study
units
represent
approximately
two
thirds
of
the
overall
water
usage
in
the
United
States
and
a
similar
proportion
of
the
population
served
by
public
water
systems.
Approximately
one
half
of
the
Nation's
land
area
is
represented
(
Leahy
and
Thompson,
1994).

To
facilitate
management
and
make
the
program
cost­
effective,
approximately
one
third
of
the
study
units
at
a
time
engage
in
intensive
assessment
for
a
period
of
3
to
5
years.
This
is
followed
by
a
period
of
less
intensive
research
and
monitoring
that
lasts
between
5
and
7
years.
This
way
all
59
study
units
rotate
through
intensive
assessment
over
a
ten­
year
period
(
Leahy
and
Thompson,
1994).
The
first
round
of
intensive
monitoring
(
1991
 
1996)
targeted
20
watersheds.
This
first
group
was
more
heavily
slanted
toward
agricultural
basins.
A
national
synthesis
of
results
from
these
study
units
and
other
research
initiatives
focusing
on
pesticides
and
nutrients
is
being
compiled
and
analyzed
(
Kolpin
et
al.,
1998;
Larson
et
al.,
1999).

For
volatile
organic
chemicals
(
VOCs),
the
national
synthesis
will
compile
data
from
the
first
and
second
rounds
of
intensive
assessments.
Study
units
assessed
in
the
second
round
represent
conditions
in
more
urbanized
basins,
but
initial
results
are
not
yet
available.
However,
VOCs
were
HCBD
 
February
2003
4­
2
analyzed
in
the
first
round
of
intensive
monitoring
and
data
are
available
for
these
study
units
(
Squillace
et
al.,
1999).
The
minimum
reporting
limit
(
MRL)
for
most
VOCs,
including
HCBD,
was
0.2
µ
g/
L
(
Squillace
et
al.,
1999).
Additional
information
on
analytical
methods
used
in
the
NAWQA
study
units,
including
method
detection
limits,
are
described
by
Gilliom
and
others
(
in
press).

Furthermore,
the
NAWQA
program
has
compiled,
by
study
unit,
data
collected
from
local,
State,
and
other
Federal
agencies
to
augment
its
own
data.
The
data
set
provides
an
assessment
of
VOCs
in
untreated
ambient
groundwater
of
the
conterminous
United
States
for
the
period
1985
 
1995
(
Squillace
et
al.,
1999).
Data
were
included
in
the
compilation
if
they
met
certain
criteria
for
collection,
analysis,
well
network
design,
and
well
construction
(
Lapham
et
al.,
1997).
They
represent
both
rural
and
urban
areas,
but
should
be
viewed
as
a
progress
report
as
NAWQA
data
continue
to
be
collected
that
may
influence
conclusions
regarding
occurrence
and
distribution
of
VOCs
(
Squillace
et
al.,
1999).

4.1.2
Results
Initial
results
published
for
the
20
NAWQA
study
units
undergoing
intensive
assessment
from
1991
 
1996
indicate
that
HCBD
was
not
detected
in
ground
water
(
Squillace
et
al.,
1999).
HCBD
also
was
not
detected
in
rural
or
urban
wells
of
the
local,
State,
and
federal
data
set
compiled
by
NAWQA.
These
data
represent
untreated
ambient
ground
water
of
the
conterminous
United
States
for
the
years
1985
 
1995
(
Squillace
et
al.,
1999).

Furthermore,
a
review
of
highway
and
urban
runoff
studies
found
no
detections
of
HCBD
(
Lopes
and
Dionne,
1998).
This
review
was
undertaken
as
part
of
the
National
Highway
Runoff
Data
and
Methodology
Synthesis
and
examined
44
studies
implemented
since
1970.

4.2
Drinking
Water
Occurrence
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1986,
required
Public
Water
Systems
(
PWSs)
to
monitor
for
specified
"
unregulated"
contaminants,
on
a
five
year
cycle,
and
to
report
the
monitoring
results
to
the
States.
Unregulated
contaminants
do
not
have
an
established
or
proposed
National
Primary
Drinking
Water
Regulation
(
NPDWR),
but
they
are
contaminants
that
were
formally
listed
and
required
for
monitoring
under
federal
regulations.
The
intent
was
to
gather
scientific
information
on
the
occurrence
of
these
contaminants
to
enable
a
decision
as
to
whether
or
not
regulations
were
needed.
All
non­
purchased
community
water
systems
(
CWSs)
and
nonpurchased
non­
transient
non­
community
water
systems
(
NTNCWSs),
with
greater
than
150
service
connections,
were
required
to
conduct
this
unregulated
contaminant
monitoring.
Smaller
systems
were
not
required
to
conduct
this
monitoring
under
federal
regulations,
but
were
required
to
be
available
to
monitor
if
the
State
decided
such
monitoring
was
necessary.
Many
States
collected
data
from
smaller
systems.
Additional
contaminants
were
added
to
the
Unregulated
Contaminant
Monitoring
(
UCM)
program
in
1991
(
U.
S.
EPA,
1991c)
for
required
monitoring
that
began
in
1993
[
57
FR
31776]
(
U.
S.
EPA,
1992c).

HCBD
has
been
monitored
under
the
SDWA
UCM
program
since
1987
[
52
FR
25720].
Monitoring
for
HCBD
under
UCM
continued
throughout
the
1990s,
but
ceased
for
small
public
HCBD
 
February
2003
4­
3
water
systems
(
PWSs)
under
a
direct
final
rule
published
on
January
8,
1999
(
64
FR
1494).
Monitoring
ended
for
large
PWSs
with
promulgation
of
the
new
Unregulated
Contaminant
Monitoring
Regulation
(
UCMR)
issued
September
17,
1999
(
64
FR
50556)
and
effective
January
1,
2001.
At
the
time
the
UCMR
lists
were
developed,
the
Agency
concluded
there
were
adequate
monitoring
data
for
a
regulatory
determination.
This
obviated
the
need
for
continued
monitoring
under
the
new
UCMR
list.

4.2.1
Data
Sources,
Data
Quality,
and
Analytical
Methods
Currently,
there
is
no
complete
national
record
of
unregulated
or
regulated
contaminants
in
drinking
water
from
public
water
systems
collected
under
SDWA.
Many
States
have
submitted
their
unregulated
contaminant
PWS
monitoring
data
to
EPA
databases,
but
there
are
issues
of
data
quality,
completeness,
and
representativeness.
Nonetheless,
a
significant
amount
of
State
data
are
available
for
UCM
contaminants
that
can
provide
estimates
of
national
occurrence.

The
National
Contaminant
Occurrence
Database
(
NCOD)
is
an
interface
to
the
actual
occurrence
data
stored
in
a
database
called
the
Safe
Drinking
Water
Information
System
(
Federal
version;
SDWIS/
FED)
and
can
be
queried
to
provide
a
summary
of
the
data
in
SDWIS/
FED
for
a
particular
contaminant.
The
data
used
in
this
report
were
derived
from
the
data
in
SDWIS/
FED
and
another
database
called
the
Unregulated
Contaminant
Information
System
(
URCIS).

The
data
in
this
report
have
been
reviewed,
edited,
and
filtered
to
meet
various
data
quality
objectives
for
the
purposes
of
this
analysis.
Hence,
not
all
data
from
a
particular
source
were
used,
only
data
meeting
the
quality
objectives
described
below.
The
sources
of
these
data,
their
quality
and
national
aggregation,
and
the
analytical
methods
used
to
estimate
a
given
contaminant's
national
occurrence
(
from
these
data)
are
discussed
in
this
section
(
for
further
details
see
U.
S.
EPA,
2001a,
b).

UCM
Rounds
1
and
2
The
1987
UCM
contaminants
include
34
volatile
organic
compounds
(
VOCs),
divided
into
two
groups:
one
with
20
VOCs
for
mandatory
monitoring,
and
the
other
with
14
VOCs
for
discretionary
monitoring
[
52
FR
25720].
HCBD
was
among
the
14
VOCs
included
for
discretionary
monitoring.
The
UCM
(
1987)
contaminants
were
first
monitored
coincident
with
the
Phase
I
regulated
contaminants,
during
the
1988
 
1992
period.
This
period
is
often
referred
to
as
"
Round
1"
monitoring.
The
monitoring
data
collected
by
the
PWSs
were
reported
to
the
States
(
as
primacy
agents),
but
there
was
no
protocol
in
place
to
report
these
data
to
EPA.
These
data
from
Round
1
were
collected
by
EPA
from
many
States
over
time.

The
Round
1
data
were
collected
in
the
URCIS.
Most
of
the
Phase
1
regulated
contaminants
were
also
VOCs.
Both
unregulated
and
regulated
VOCs
are
analyzed
using
the
same
sample
and
the
same
laboratory
methods.
Hence,
the
URCIS
database
includes
data
on
all
of
these
62
contaminants:
the
34
UCM
(
1987)
VOCs;
the
21
regulated
Phase
1
VOCs;
2
regulated
synthetic
organic
contaminants
(
SOCs);
and
5
miscellaneous
contaminants
that
were
voluntarily
reported
by
some
States
(
e.
g.,
isomers
of
other
organic
contaminants).
HCBD
 
February
2003
4­
4
The
1993
UCM
contaminants
include
13
SOCs
and
1
inorganic
contaminant
(
IOC)
[
56
FR
3526].
Monitoring
for
the
UCM
(
1993)
contaminants
began
coincident
with
the
Phase
II/
V
regulated
contaminants
in
1993
through
1998.
This
is
often
referred
to
as
"
Round
2"
monitoring.
The
UCM
(
1987)
contaminants
were
also
included
in
the
Round
2
monitoring.
As
with
other
monitoring
data,
PWSs
reported
these
results
to
the
States.
EPA,
during
the
past
several
years,
requested
that
the
States
submit
these
historic
data
to
EPA.

The
details
of
the
actual
individual
monitoring
periods
are
complex.
The
timing
of
required
monitoring
was
staggered
related
to
different
size
classes
of
PWSs,
and
the
program
was
implemented
somewhat
differently
by
different
States.
Round
1
includes
the
period
from
1988
 
1992,
it
also
includes
results
from
samples
analyzed
prior
to
1988
(
for
further
details
see
U.
S.
EPA,
2001a,
b).

Developing
a
Nationally
Representative
Perspective
The
URCIS
and
SDWIS/
FED
databases
contain
contaminant
occurrence
data
from
a
total
of
40
and
35
primacy
entities
(
largely
States),
respectively.
However,
data
from
some
States
are
incomplete
and
biased.
Furthermore,
the
national
representativeness
of
the
data
is
questionable
because
the
data
were
not
collected
in
a
systematic
or
random
statistical
framework.
These
State
data
could
be
heavily
skewed
to
low­
occurrence
or
high­
occurrence
settings.
Hence,
the
data
were
evaluated
based
on
pollution­
potential
indicators
and
the
spatial/
hydrologic
diversity
of
the
nation.
This
evaluation
enabled
the
construction
of
a
cross­
section
from
the
available
State
data
sets
that
provides
a
reasonable
representation
of
national
occurrence.

A
national
cross­
section
from
State
SDWA
contaminant
databases
was
established
using
the
approach
developed
for
the
EPA
report
A
Review
of
Contaminant
Occurrence
in
Public
Water
Systems
(
U.
S.
EPA,
1999c).
This
approach
was
developed
to
support
occurrence
analyses
for
EPA's
Chemical
Monitoring
Reform
(
CMR)
evaluation.
It
was
supported
by
peer
reviewers
and
stakeholders
because
it
is
clear,
repeatable,
and
understandable.
The
approach
cannot
provide
a
"
statistically
representative"
sample
because
the
original
monitoring
data
were
not
collected
or
reported
in
an
appropriate
fashion.
However,
the
resultant
"
national
cross­
section"
of
States
should
provide
a
clear
indication
of
the
central
tendency
of
the
national
data.
The
remainder
of
this
section
provides
a
summary
description
of
how
the
national
cross­
sections
for
the
URCIS
(
Round
1)
and
SDWIS/
FED
(
Round
2)
databases
were
developed.
The
details
of
the
approach
are
presented
in
U.
S.
EPA
(
2001a,
b).

Cross­
Section
Development
As
a
first
step
in
developing
the
cross­
section,
the
State
data
contained
in
the
URCIS
database
(
which
contains
Round
1
monitoring
results)
and
SDWIS/
FED
database
(
which
contains
Round
2
monitoring
results)
were
evaluated
for
completeness
and
quality.
For
both
the
URCIS
(
Round
1)
and
SDWIS/
FED
(
Round
2)
databases,
some
State
data
were
unusable
for
a
variety
of
reasons.
Some
States
reported
only
detections,
or
their
data
had
incorrect
units.
Datasets
only
including
detections
are
obviously
biased.
Other
problems
included
incomplete
data
sets
without
all
PWSs
reporting.
Also,
data
from
Washington,
D.
C.
and
the
Virgin
Islands
were
excluded
from
this
HCBD
 
February
2003
4­
5
analysis
because
it
was
difficult
to
evaluate
them
for
the
current
purposes
in
relation
to
complete
State
data
(
U.
S.
EPA,
2001a,
Sections
II
and
III).

The
balance
of
the
States
remaining
after
the
data
quality
screening
were
then
examined
to
establish
a
national
cross­
section.
This
step
was
based
on
evaluating
the
States'
pollution
potential
and
geographic
coverage
in
relation
to
all
States.
Pollution
potential
is
considered
to
ensure
a
selection
of
States
that
represent
the
range
of
likely
contaminant
occurrence
and
a
balance
with
regard
to
likely
high
and
low
occurrence.
Geographic
consideration
is
included
so
that
the
wide
range
of
climatic
and
hydrogeologic
conditions
across
the
United
States
are
represented,
again
balancing
the
varied
conditions
that
affect
transport
and
fate
of
contaminants
(
U.
S.
EPA,
2001b,
Sections
III.
A.
and
III.
B.).

The
cross­
section
States
were
selected
to
represent
a
variety
of
pollution
potential
conditions.
Two
primary
pollution
potential
indicators
were
used.
The
first
factor
selected
indicates
pollution
potential
from
manufacturing/
population
density
and
serves
as
an
indicator
of
the
potential
for
VOC
contamination
within
a
State.
Agriculture
was
selected
as
the
second
pollution
potential
indicator
because
the
majority
of
SOCs
of
concern
are
pesticides
(
U.
S.
EPA,
2001b,
Section
III.
A.).
The
50
individual
States
were
ranked
from
highest
to
lowest
based
on
the
pollution
potential
indicator
data.
For
example,
the
State
with
the
highest
ranking
for
pollution
potential
from
manufacturing
received
a
ranking
of
1
for
this
factor
and
the
State
with
the
lowest
value
was
ranked
as
number
50.
States
were
ranked
for
their
agricultural
chemical
use
status
in
a
similar
fashion.

The
States'
pollution
potential
rankings
for
each
factor
were
subdivided
into
four
quartiles
(
from
highest
to
lowest
pollution
potential).
The
cross­
section
States
were
chosen
from
all
quartiles
for
both
pollution
potential
factors
to
ensure
representation,
as
follows:
States
with
high
agrochemical
pollution
potential
rankings
and
high
manufacturing
pollution
potential
rankings;
States
with
high
agrochemical
pollution
potential
rankings
and
low
manufacturing
pollution
potential
rankings;
States
with
low
agrochemical
pollution
potential
rankings
and
high
manufacturing
pollution
potential
rankings;
and
States
with
low
agrochemical
pollution
potential
rankings
and
low
manufacturing
pollution
potential
rankings
(
U.
S.
EPA,
2001b,
Section
III.
B.).
In
addition,
some
secondary
pollution
potential
indicators
were
considered
to
further
ensure
that
the
cross­
section
States
included
the
spectrum
of
pollution
potential
conditions
(
high
to
low).

The
data
quality
screening,
pollution
potential
rankings,
and
geographic
coverage
analysis
established
national
cross­
sections
of
24
Round
1
(
URCIS)
States
and
20
Round
2
(
SDWIS/
FED)
States.
In
each
cross­
section,
the
States
provide
good
representation
of
the
Nation's
varied
climatic
and
hydrogeologic
regimes
and
the
breadth
of
pollution
potential
for
the
contaminant
groups
(
Table
4­
1
and
Figure
4­
1).

Cross­
Section
Evaluation
To
evaluate
and
validate
the
method
for
creating
the
national
cross­
sections,
the
method
was
used
to
create
smaller
State
subsets
from
the
24­
State,
Round
1
cross­
section
and
aggregations.
Again,
States
were
chosen
to
achieve
a
balance
from
the
quartiles
describing
pollution
potential,
and
a
balanced
geographic
distribution,
to
incrementally
build
HCBD
 
February
2003
4­
6
Table
4­
1.
Cross­
section
States
for
Round
1
(
24
States)
and
Round
2
(
20
States).

Round
1
(
URCIS)
Round
2
(
SDWIS/
FED)

Alabama
Alaska*
Arizona
California
Florida
Georgia
Hawaii
Illinois
Indiana
Iowa
Kentucky*
Maryland*
Minnesota*
Montana
New
Jersey
New
Mexico*
North
Carolina*
Ohio*
South
Dakota
Tennessee
Utah
Washington*
West
Virginia
Wyoming
Alaska*
Arkansas
Colorado
Kentucky*
Maine
Maryland*
Massachusetts
Michigan
Minnesota*
Missouri
New
Hampshire
New
Mexico*
North
Carolina*
North
Dakota
Ohio*
Oklahoma
Oregon
Rhode
Island
Texas
Washington*

*
cross­
section
State
in
both
Round
1
and
Round
2
Figure
4­
1.
Geographic
Distribution
of
Cross­
section
States
for
Round
1
(
left)
and
Round
2
(
right).

subset
cross­
sections
of
various
sizes.
For
example,
the
Round
1
cross­
section
was
tested
with
subsets
of
4,
8
(
the
first
4­
State
subset
plus
4
more
States),
and
13
(
8­
State
subset
plus
5)
States.
Two
additional
cross­
sections
were
included
in
the
analysis
for
comparison:
a
cross­
section
composed
of
the
16
biased
States
eliminated
from
the
24­
State
cross­
section
for
data
quality
reasons
and
a
crosssection
composed
of
all
40
Round
1
States
(
U.
S.
EPA,
2001,
Section
III.
B.
1).
HCBD
 
February
2003
4­
7
These
Round
1
incremental
cross­
sections
were
then
used
to
evaluate
occurrence
for
an
array
of
both
high
and
low
occurrence
contaminants.
The
comparative
results
illustrate
several
points.
The
results
are
quite
stable
and
consistent
for
the
8­,
13­
and
24­
State
cross­
sections.
They
are
much
less
so
for
the
4­
State,
16­
State
(
biased),
and
40­
State
(
all
Round
1
States)
cross­
sections.
The
4­
State
cross­
section
is
apparently
too
small
to
provide
balance
both
geographically
and
with
pollution
potential,
a
finding
that
concurs
with
past
work
(
U.
S.
EPA,
1999c).
The
CMR
analysis
suggested
that
a
minimum
of
6
 
7
States
was
needed
to
provide
balance
both
geographically
and
with
pollution
potential,
and
the
CMR
report
used
8
States
out
of
the
available
data
for
its
nationally
representative
cross­
section.
The
16­
State
and
40­
State
cross­
sections,
both
including
the
biased
States,
provided
occurrence
results
that
were
unstable
and
inconsistent
for
a
variety
of
reasons
associated
with
their
data
quality
problems
(
U.
S.
EPA,
2001,
Section
III.
B.
1).

The
8­,
13­,
and
24­
State
cross­
sections
provide
very
comparable
results,
are
consistent,
and
are
usable
as
national
cross­
sections
to
provide
estimates
of
contaminant
occurrence.
Including
greater
data
from
more
States
improves
the
national
representation
and
the
confidence
in
the
results,
as
long
as
the
States
are
balanced
related
to
pollution
potential
and
spatial
coverage.
The
24­
and
20­
State
cross­
sections
provide
the
best,
nationally
representative
cross­
sections
for
the
Round
1
and
Round
2
data.

Data
Management
and
Analysis
The
cross­
section
analyses
focused
on
occurrence
at
the
water
system
level;
i.
e.,
the
summary
data
presented
discuss
the
percentage
of
public
water
systems
with
detections,
not
the
percentage
of
samples
with
detections.
By
normalizing
the
analytical
data
to
the
system
level,
skewness
inherent
in
the
sample
data,
particularly
over
the
multi­
year
period
covered
in
the
URCIS
data,
is
avoided.
System
level
analysis
was
used
since
a
PWS
with
a
known
contaminant
problem
usually
has
to
sample
more
frequently
than
a
PWS
that
has
never
detected
the
contaminant.
Obviously,
the
results
of
a
simple
computation
of
the
percentage
of
samples
with
detections
(
or
other
statistics)
can
be
skewed
by
the
more
frequent
sampling
results
reported
by
the
contaminated
site.
This
level
of
analysis
is
conservative.
For
example,
a
system
need
only
have
a
single
sample
with
an
analytical
result
greater
than
the
MRL,
i.
e.,
a
detection,
to
be
counted
as
a
system
with
a
result
"
greater
than
the
MRL."

Also,
the
data
used
in
the
analyses
were
limited
to
only
those
data
with
confirmed
water
source
and
sampling
type
information.
Only
standard
SDWA
compliance
samples
were
used;
"
special"
samples,
or
"
investigation"
samples
(
investigating
a
contaminant
problem
that
would
bias
results),
or
samples
of
unknown
type
were
not
used
in
the
analyses.
Various
quality
control
and
review
checks
were
made
of
the
results,
including
follow­
up
questions
to
the
States
providing
the
data.
Many
of
the
most
intractable
data
quality
problems
encountered
occurred
with
older
data.
These
problematic
data
were,
in
some
cases,
simply
eliminated
from
the
analysis.
For
example,
when
the
number
of
data
with
problems
were
insignificant
relative
to
the
total
number
of
observations,
they
were
dropped
from
the
analysis
(
For
further
details,
see
Cadmus,
2000).
HCBD
 
February
2003
4­
8
Occurrence
Analysis
To
evaluate
national
contaminant
occurrence,
a
two­
stage
analytical
approach
has
been
developed.
The
first
stage
of
analysis
provides
a
straight­
forward,
conservative,
broad
evaluation
of
occurrence
of
the
Contaminant
Candidate
List
(
CCL)
preliminary
regulatory
determination
priority
contaminants
as
described
above.
These
descriptive
statistics
are
summarized
here.
Based
on
the
findings
of
the
Stage
1
Analysis,
EPA
will
determine
whether
more
intensive
statistical
evaluations,
the
Stage
2
Analysis,
may
be
warranted
to
generate
national
probability
estimates
of
contaminant
occurrence
and
exposure
for
priority
contaminants
(
for
details
on
this
two
stage
analytical
approach
see
Cadmus,
2000)

The
summary
descriptive
statistics
presented
in
Table
4­
2
for
HCBD
are
a
result
of
the
Stage
1
analysis
and
include
data
from
both
Round
1
(
URCIS,
1987
 
1992)
and
Round
2
(
SDWIS/
FED,
1993
 
1997)
cross­
section
States.
Included
are
the
total
number
of
samples,
the
percent
samples
with
detections,
the
99th
percentile
concentration
of
all
samples,
the
99th
percentile
concentration
of
samples
with
detections,
and
the
median
concentration
of
samples
with
detections.
The
percentages
of
PWSs
and
population
served
indicate
the
proportion
of
PWSs
whose
analytical
results
showed
a
detection(
s)
of
the
contaminant
(
simple
detection,
>
MRL)
at
any
time
during
the
monitoring
period;
or
a
detection(
s)
greater
than
half
the
Health
Reference
Level
(
HRL);
or
a
detection(
s)
greater
than
the
Health
Reference
Level.
The
Health
Reference
Level,
0.9
µ
g/
L,
is
a
preliminary
estimated
health
effect
level
used
for
this
analysis.
The
HRL
was
derived
using
the
10­
6
cancer
risk
as
calculated
by
the
linear
method
using
a
body
weight
to
the
three
quarter
power
(
section
8.8.2;
slope
factor
4
x
10­
2
(
mg/
kg/
day)­
1.

When
monitoring
results
were
compared
to
a
value
of
one­
half
of
the
HRL,
0.16%
of
Round
1
(
106
systems)
and
0.08%
of
Round
2
(
51
systems)
water
supplies
exceeded
this
benchmark
at
least
once
during
the
reporting
period.
The
percentages
of
water
supplies
that
exceeded
the
HRL
at
least
once
in
Round
1
and
Round
2
monitoring
were
0.11%
(
74
systems)
and
0.02%
(
11
systems),
respectively.

The
99th
percentile
concentration
is
used
here
as
a
summary
statistic
to
indicate
the
upper
bound
of
occurrence
values
because
maximum
values
can
be
extreme
values
(
outliers)
that
sometimes
result
from
sampling
or
reporting
error.
The
99th
percentile
concentration
is
presented
for
both
the
samples
with
only
detections
and
all
of
the
samples
because
the
value
for
the
99th
percentile
concentration
of
all
samples
is
below
the
MRL
(
denoted
by
"<"
in
Table
4­
2).
For
the
same
reason,
summary
statistics
such
as
the
95th
percentile
concentration
of
all
samples
or
the
median
(
or
mean)
concentration
of
all
samples
are
omitted
because
these
also
are
all
"<"
values.
This
is
the
case
because
only
0.1
to
0.05%
of
all
samples
recorded
detections
of
HCBD
in
Round
1
and
Round
2.

As
a
convention,
a
value
of
half
the
MRL
is
often
used
as
an
estimate
of
the
concentration
of
a
contaminant
in
samples/
systems
whose
results
are
less
than
the
MRL.
With
a
contaminant
with
relatively
low
occurrence
such
as
HCBD
in
drinking
water
occurrence
databases,
the
median
or
mean
value
of
occurrence
using
this
assumption
would
be
half
the
MRL
(
0.5
×
MRL).
However,
for
these
occurrence
data
this
is
not
straightforward.
For
Round
1
and
Round
2,
States
have
reported
a
wide
HCBD
 
February
2003
4­
9
range
of
values
for
the
MRLs.
This
is
in
part
related
to
State
data
management
differences
as
well
as
real
differences
in
analytical
methods,
laboratories,
and
other
factors.

The
situation
can
cause
confusion
when
examining
descriptive
statistics
for
occurrence.
For
example,
the
modal
MRL
value
for
the
Round
1
samples
is
0.50
µ
g/
L
 
a
value
twice
as
large
as
the
median
concentration
of
detections
for
Round
1
(
0.25
µ
g/
L)
(
This
occurs
because
some
States
and/
or
systems
reporting
detections
were
using
a
lower
MRL
and
had
positive
results
lower
than
the
MRL
used
by
other
States
or
systems).
For
Round
2,
most
States
reported
non­
detections
as
zeros
resulting
in
a
modal
MRL
value
of
zero.
By
definition
the
MRL
cannot
be
zero.
This
is
an
artifact
of
State
data
management
systems.
Because
a
simple
meaningful
summary
statistic
is
not
available
to
describe
the
various
reported
MRLs,
and
to
avoid
confusion,
MRLs
are
not
reported
in
the
summary
table,
but
rather
are
designated
as
"
variable"
(
Table
4­
2).

In
Table
4­
2,
national
occurrence
is
estimated
by
extrapolating
the
summary
statistics
for
the
24­
and
20­
State
cross­
sections
to
national
numbers
for
systems,
and
population
served
by
systems,
from
the
Water
Industry
Baseline
Handbook,
Second
Edition
(
U.
S.
EPA,
2000e).
From
the
handbook,
the
total
number
of
community
water
systems
(
CWSs)
plus
non­
transient,
non­
community
water
systems
(
NTNCWSs)
is
65,030,
and
the
total
population
served
by
CWSs
plus
NTNCWSs
is
213,008,182
persons
(
see
Table
4­
2).
To
arrive
at
the
national
occurrence
estimate
for
a
particular
cross­
section,
the
national
estimate
for
PWSs
(
or
population
served
by
PWSs)
is
simply
multiplied
by
the
percentage
for
the
given
summary
statistic.
[
i.
e.,
for
Round
1,
the
national
estimate
for
the
total
number
of
PWSs
with
detections
(
228)
is
the
product
of
the
percentage
of
Round
1
PWSs
with
detections
(
0.35%)
and
the
national
estimate
for
the
total
number
of
PWSs
(
65,030)].

Because
the
State
data
used
for
the
cross­
section
are
not
a
strict
statistical
sample,
national
extrapolations
of
these
Stage
1
analytical
results
can
be
problematic,
especially
for
contaminants
with
very
low
occurrence
like
hexachlorobutadiene
and
other
CCL
regulatory
determination
priority
contaminants.
For
this
reason,
the
nationally
extrapolated
estimates
of
occurrence
based
on
Stage
1
results
are
not
presented
in
the
CCL
Federal
Register
Notice.
The
presentation
in
the
Federal
Register
Notice
of
only
the
actual
results
of
the
cross­
section
analysis
maintains
a
straight­
forward
presentation,
and
the
integrity
of
the
data,
for
stakeholder
review.
The
nationally
extrapolated
Stage
1
occurrence
values
are
presented
here,
however,
to
provide
additional
perspective.
A
more
rigorous
statistical
modeling
effort,
the
Stage
2
analysis,
could
be
conducted
on
the
cross­
section
data
(
Cadmus,
2001).
The
Stage
2
results
would
be
more
statistically
robust
and
more
suitable
to
national
extrapolation.
This
approach
would
provide
a
probability
estimate
and
would
also
allow
for
better
quantification
of
estimation
error.

Round
1(
1987
 
1992)
and
Round
2
(
1993
 
1997)
data
were
not
merged
because
they
represent
different
time
periods,
different
States
(
only
eight
States
are
represented
in
both
rounds),
and
each
round
has
different
data
management
and
data
quality
problems.
The
two
rounds
are
only
merged
for
the
simple
spatial
analysis
overview
presented
in
Section
4.2
and
Figures
4­
2
and
4­
4.
HCBD
 
February
2003
4­
10
4.2.2
Results
Occurrence
Estimates
While
States
with
detections
of
HCBD
are
widespread
(
Figure
4­
2),
the
percentages
of
PWSs
by
State
with
detections
are
low
(
Table
4­
2).
In
aggregate,
the
cross­
sections
show
only
0.2
 
0.4
%
of
PWSs
in
both
rounds
experienced
detections
(>
MRL),
affecting
0.9
 
2.4%
of
the
population
served
(
approximately
2
 
5
million
people).
Percentages
of
PWSs
with
detections
greater
than
half
the
Health
Reference
Level
(>
½
HRL)
are
slightly
lower:
0.1
 
0.2%.
The
percentage
of
PWSs
exceeding
the
Health
Reference
Level
(>
HRL)
for
both
rounds
is
very
small
(
see
also
Figure
4­
4).
The
percentage
of
PWSs
that
experienced
detections
>
HRL
in
Rounds
1
and
2
are
0.1%
and
0.02%,
respectively;
affecting
a
population
of
approximately
780,000
and
10,000,
respectively.

There
are
some
qualifying
notes
for
both
rounds
of
data
that
warrant
discussion.
The
Round
1
estimates
of
PWSs
affected
by
HCBD
are
influenced
by
the
State
of
Florida
(
Table
4­
2;
Figures
4­
3
and
4­
4).
This
State
reports
that
5.4%
of
its
PWSs
experienced
detections
greater
than
the
HRL
during
Round
1,
a
value
considerably
greater
than
the
next
highest
State
(
1.5%).
This
suggests
that
Florida's
data
for
HCBD
is
incomplete
and
may
be
biased.
Out
of
855
Florida
PWSs
reporting
contaminant
data
for
Round
1
monitoring,
only
112
provided
data
for
HCBD
(
U.
S.
EPA,
2001a).
Also,
the
5.4%
of
systems
reporting
detections
all
reported
concentrations
greater
than
the
Health
Reference
Level.
These
figures
suggest
that
perhaps
only
systems
experiencing
problems
submitted
data
for
HCBD,
biasing
Florida's
results
for
occurrence
measures
examined
in
this
report.

The
large
values
for
the
Round
2
national
estimates
of
population
served
with
detections
greater
than
the
MRL
and
greater
than
half
the
HRL
are
influenced
by
the
inclusion
of
one
PWS
serving
a
very
large
population
(
1.5
million
people).
While
the
percentage
of
systems
with
detections
of
HCBD
are
similar
(
both
rounds
show
low
values,
0.2
 
0.4%
PWSs
>
MRL),
the
difference
in
population
served
results
in
a
larger
difference
in
the
population
extrapolations.

Note
that
for
the
Round
1
cross­
section,
the
total
number
of
PWSs
(
and
the
total
population
served
by
the
PWSs)
is
not
the
sum
of
the
number
of
ground
water
and
surface
water
systems
(
or
the
populations
served
by
those
systems).
Because
some
public
water
systems
are
seasonally
classified
as
either
surface
or
ground
water,
some
systems
may
be
counted
in
both
categories.
The
population
numbers
for
the
Round
1
cross­
section
are
also
incomplete.
Not
all
of
the
PWSs
for
which
occurrence
data
was
submitted
reported
the
population
they
served.
However,
the
population
numbers
presented
in
Table
4­
2
for
the
Round
1
cross­
section
are
reported
from
94%
of
the
systems.

The
national
estimates
extrapolated
from
Round
1
and
Round
2
PWS,
numbers
and
populations
are
not
additive.
In
addition
to
the
Round
1
classification
and
reporting
issues
outlined
above,
the
proportions
of
surface
water
and
ground
water
PWSs,
and
populations
served
by
them,
are
different
between
the
Round
1
and
2
cross­
sections
and
the
national
estimates.
For
example,
approximately
48%
of
the
population
served
by
PWSs
in
the
Round
1
cross­
section
States
are
served
by
surface
water
PWSs
(
Table
4­
2).
Nationally,
however,
that
proportion
changes
to
60%.
HCBD
 
February
2003
4­
11
Table
4­
2.
Summary
Occurrence
Statistics
for
Hexachlorobutadiene.

Frequency
Factors
24­
State
Cross­
Section1
(
Round
1)
20­
State
Cross­
Section2
(
Round
2)
National
System
&
Population
Numbers3
Total
Number
of
Samples
42,839
93,585
­­

Percent
of
Samples
with
Detections
0.13%
0.05%
­­

99th
Percentile
Concentration
(
all
samples)
<
(
Non­
detect)
<
(
Non­
detect)
­­

Health
Reference
Level
0.9
:
g/
L
0.9
:
g/
L
­­

Minimum
Reporting
Level
(
MRL)
Variable*
Variable*
­­

99th
Percentile
Concentration
of
Detections
10
:
g/
L
1.5
:
g/
L
­­

Median
Concentration
of
Detections
0.25
:
g/
L
0.30
:
g/
L
­­

Total
Number
of
PWSs
12,284
22,736
65,030
Number
of
GW
PWSs
10,980
20,380
59,440
Number
of
SW
PWSs
1,385
2,356
5,590
Total
Population
71,582,571
67,075,493
213,008,182
Population
of
GW
PWSs
40,399,177
24,960,222
85,681,696
Population
of
SW
PWSs
34,418,834
42,115,271
127,326,486
National
Extrapolation4
Occurrence
by
System
Round
1
Round
2
PWSs
with
detections
(>
MRL)
0.350%
0.180%
228
117
Range
of
Cross­
Section
States
0
 
5.36%
0
 
3.36%
N/
A
N/
A
GW
PWSs
with
detections
0.301%
0.132%
179
79
SW
PWSs
with
detections
0.722%
0.594%
40
33
PWSs
>
½
Health
Reference
Level
(
HRL)
0.163%
0.079%
106
51
Range
of
Cross­
Section
States
0
 
5.36%
0
 
0.51%
N/
A
N/
A
GW
PWSs
>
½
Health
Reference
Level
0.118%
0.064%
70
38
SW
PWSs
>
½
Health
Reference
Level
0.505%
0.212%
28
12
PWSs
>
Health
Reference
Level
0.114%
0.018%
74
11
Range
of
Cross­
Section
States
0
 
5.36%
0
 
0.24%
N/
A
N/
A
GW
PWSs
>
Health
Reference
Level
0.064%
0.005%
38
3
SW
PWSs
>
Health
Reference
Level
0.505%
0.127%
28
7
Occurrence
by
Population
Served
Round
1
Round
2
PWS
Population
Served
with
detections
0.896%
2.360%
1,909,000
5,027,000
Table
4­
2
(
continued)

Frequency
Factors
24­
State
Cross­
Section1
(
Round
1)
20­
State
Cross­
Section2
(
Round
2)
National
System
&
Population
Numbers3
HCBD
 
February
2003
4­
12
Range
of
Cross­
Section
States
0
 
11.38%
0
 
29.93%
N/
A
N/
A
GW
PWS
Population
with
detections
1.458%
0.186%
1,249,000
159,000
SW
PWS
Population
with
detections
0.153%
3.649%
194,000
4,646,000
PWS
Population
Served
>
½
Health
Ref
Level
0.569%
2.331%
1,213,000
4,965,000
Range
of
Cross­
Section
States
0
 
11.38%
0
 
29.92%
N/
A
N/
A
GW
PWS
Population
>
½
Health
Ref
Level
0.978%
0.177%
838,000
152,000
SW
PWS
Population
>
½
Health
Ref
Level
0.036%
3.607%
46,000
4,593,000
PWS
Population
Served
>
Health
Ref
Level
0.367%
0.005%
781,000
10,000
Range
of
Cross­
Section
States
0
 
9.66%
0
 
0.02%
N/
A
N/
A
GW
PWS
Population
>
Health
Ref
Level
0.619%
0.011%
531,000
9,000
SW
PWS
Population
>
Health
Ref
Level
0.036%
0.001%
46,000
1,000
1.
Summary
Results
based
on
data
from
24­
State
Cross­
Section,
from
URCIS,
UCM
(
1987)
Round
1.
2.
Summary
Results
based
on
data
from
20­
State
Cross­
Section,
from
SDWIS/
FED,
UCM
(
1993)
Round
2.
3.
Total
PWS
and
population
numbers
are
from
EPA
March
2000
Water
Industry
Baseline
Handbook.
4.
National
extrapolations
are
from
the
24­
State
data
and
20­
State
data
using
the
Baseline
Handbook
system
and
population
numbers.
*
see
text
for
discussion
­
PWS
=
Public
Water
Systems;
GW
=
Ground
Water;
SW
=
Surface
Water;
MRL
=
Minimum
Reporting
Level
(
for
laboratory
analyses);
­
Health
Reference
Level
=
Health
Reference
Level,
an
estimated
health
effect
level
used
for
preliminary
assessment
for
this
review;
N/
A
=
Not
Applicable
­
The
Health
Reference
Level
used
for
hexachlorobutadiene
is
0.9
:
g/
L.
This
is
a
draft
value
for
working
review
only.
­
Total
Number
of
Samples
=
the
total
number
of
analytical
records
for
hexachlorobutadiene.
­
99th
Percentile
Concentration
=
the
concentration
value
of
the
99th
percentile
of
either
all
analytical
results
or
just
the
samples
with
detections
(
in
:
g/
L).
­
Median
Concentration
of
Detections
=
the
median
analytical
value
of
all
the
detections
(
analytical
results
greater
than
the
MRL)
(
in
:
g/
L).
­
Total
Number
of
PWSs
=
the
total
number
of
public
water
systems
with
records
for
hexachlorobutadiene.
­
Total
Population
Served
=
the
total
population
served
by
public
water
systems
with
records
for
hexachlorobutadiene.
­
%
PWS
with
detections,
%
PWS
>
½
Health
Reference
Level,
%
PWS
>
Health
Reference
Level
=
percent
of
the
total
number
of
public
water
systems
with
at
least
one
analytical
result
that
exceeded
the
MRL,
½
Health
Reference
Level,
Health
Reference
Level,
respectively.
­
%
PWS
Population
Served
with
detections,
%
PWS
Population
Served
>
½
Health
Reference
Level,
%
PWS
Population
Served
>
Health
Reference
Level
=
percent
of
the
total
population
served
by
PWSs
with
at
least
one
analytical
result
exceeding
the
MRL,
½
Health
Reference
Level,
or
the
Health
Reference
Level,
respectively.

Both
Round
1
and
Round
2
national
cross­
sections
show
a
proportionate
balance
in
source
waters.
Nationally,
91%
of
PWSs
use
ground
water
(
and
9%
surface
waters):
Round
1
shows
89%,
and
Round
2
shows
90%
of
systems
using
ground
water.
The
relative
populations
served
are
not
as
closely
comparable.
Nationally,
about
40%
of
the
population
is
served
by
PWSs
using
groundwater
(
and
60%
by
surface
water).
Round
2
data
is
most
representative
with
37%
of
the
cross­
section
population
served
by
ground
water;
Round
1
shows
about
55%.

There
are
differences
in
the
occurrence
results
between
Round
1
and
Round
2,
as
should
be
expected.
The
differences
are
not
great,
however,
particularly
when
comparing
the
proportions
of
HCBD
 
February
2003
4­
13
systems
affected.
The
results
range
from
0.2
 
0.4%
of
PWSs
with
detections
of
HCBD
and
range
from
0.02
 
0.1%
of
PWSs
with
detections
greater
than
the
Health
Reference
Level
of
0.9
µ
g/
L.
These
are
not
substantively
different,
given
the
data
sources.

The
differences
in
the
population
extrapolations
appear
greater,
but
still
constitute
relatively
small
proportions
of
the
population.
The
most
pronounced
difference
is
in
the
estimate
of
the
population
served
by
PWSs
with
detections
greater
than
the
Health
Reference
Level,
ranging
from
10,000
to
780,000.
In
both
cases,
this
is
less
than
0.5%
of
the
population.
The
difference
in
this
category
is
largely
driven
by
the
Florida
data
in
Round
1,
as
discussed
above.

The
Round
2
cross­
section
provides
a
better
proportional
balance
related
to
the
national
population
of
PWSs
and
may
have
fewer
reporting
problems
than
Round
1
(
i.
e.,
incomplete
population
numbers,
Florida).
The
larger
estimate
of
the
national
population
served
by
PWSs
with
detections
greater
than
the
Health
Reference
Level
using
Round
1
data
can
also
provide
an
upper
bound
estimate
in
considering
the
data.

Regional
Patterns
Occurrence
results
are
displayed
graphically
by
State
in
Figures
4­
2,
4­
3,
and
4­
4
to
assess
whether
any
distinct
regional
patterns
of
occurrence
are
present.
Combining
Round
1
and
Round
2
data
(
Figure
4­
2),
there
are
47
States
reporting.
Six
of
those
States
have
no
data
for
HCBD,
while
another
21
have
no
detections
of
the
chemical.
The
remaining
20
States
have
detected
HCBD
in
drinking
water
and
are
well
distributed
throughout
the
United
States.

The
simple
spatial
analysis
presented
in
Figures
4­
2,
4­
3,
and
4­
4
suggests
that
special
regional
analyses
are
not
warranted.
Florida's
possible
bias
is
notable,
however.
While
no
clear
geographical
patterns
of
occurrence
are
apparent,
comparisons
with
environmental
use
and
release
information
are
useful
(
see
also
Section
2.2).
Five
of
the
eight
Toxic
Release
Inventory
States
that
reported
releases
of
HCBD
into
the
environment
between
1988
and
1998
have
also
detected
the
chemical
in
PWS
sampling.
Of
the
remaining
three
(
Kansas,
Louisiana,
and
California),
Kansas
hasn't
reported
any
data
for
either
Round
1
or
2.
Also,
of
the
eight
States
with
detections
of
HCBD
at
CERCLA
National
Priorities
List
(
NPL)
hazardous
waste
sites,
five
have
detected
the
chemical
in
drinking
water.
Finally,
six
of
the
States
detecting
HCBD
in
PWS
samples
have
also
detected
it
in
site
samples
reported
to
the
ATSDR's
HazDat
database.
It
is
interesting
to
note
that
neither
Alabama
nor
Florida,
the
two
States
with
the
highest
percentage
of
PWSs
with
detections
greater
than
the
Health
Reference
level,
are
Toxic
Release
Inventory
(
TRI)
States
for
HCBD,
nor
do
they
have
CERCLA
NPL
sites
with
detections
of
the
chemical
(
Figure
4­
4).
HCBD
 
February
2003
4­
14
All
States
Hexachlorobutadiene
Detections
in
Round
1
and
Round
2
States
not
in
Round
1
or
Round
2
No
data
for
Hexachlorobutadiene
States
with
No
Detections
(
No
PWSs
>
MRL)
States
with
Detections
(
Any
PWSs
>
MRL)
Figure
4­
2.
States
with
PWSs
with
Detections
of
Hexachlorobutadiene
for
all
States
with
Data
in
URCIS
(
Round
1)
and
SDWIS/
FED
(
Round
2).
HCBD
 
February
2003
4­
15
*
State
of
Florida
is
an
outlier
with
5.36%
PWS
>
MRL
Hexachlorobutadiene
Occurrence
in
Round
1
States
not
in
Cross­
Section
No
data
for
Hexachlorbutadiene
0.00%
PWSs
>
MRL
0.01
­
1.00%
PWSs
>
MRL
1.00
­
3.50%
PWSs
>
MRL*

Hexachlorobutadiene
Occurrence
in
Round
2
States
not
in
Cross­
Section
No
data
for
Hexachlorbutadiene
0.00%
PWSs
>
MRL
0.01
­
1.00%
PWSs
>
MRL
1.00
­
3.50%
PWSs
>
MRL
Figure
4­
3.
States
with
PWSs
with
Detections
of
Hexachlorobutadiene
(
any
PWSs
with
results
greater
than
the
Minimum
Reporting
Level
[
MRL])
for
Round
1
(
above)
and
Round
2
(
below)
Cross­
section
States.
HCBD
 
February
2003
4­
16
*
State
of
Florida
is
an
outlier
with
5.36%
PWS
>
MRL
Hexachlorobutadiene
Occurrence
in
Round
1
and
Round
2
States
not
in
Cross­
Section
No
data
for
Hexachlorbutadiene
0.00%
PWSs
>
MRL
0.01
­
1.00%
PWSs
>
MRL
1.00
­
3.50%
PWSs
>
MRL*

Hexachlorobutadiene
Occurrence
in
Round
1
and
Round
2
States
not
in
Cross­
Section
No
data
for
Hexachlorobutadiene
0.00%
PWSs
>
HRL
0.01
­
1.00%
PWSs
>
HRL
1.00
­
3.50%
PWSs
>
HRL*
*
State
of
Florida
is
an
outlier
with
5.36%
PWS
>
HRL
Figure
4­
4.
Cross­
section
States
(
Round
1
and
Round
2
combined)
with
PWSs
with
Detections
of
Hexachlorobutadiene
(
above)
and
concentrations
greater
than
the
Health
Reference
Level
(
HRL;
below).
HCBD
 
February
2003
4­
17
4.3
Conclusions
While
there
have
not
been
detections
of
the
chemical
in
ambient
water
reported
in
USGS
NAWQA
studies
to
date,
hexachlorobutadiene
has
been
detected
at
a
very
low
percentage
of
ATSDR
HazDat
sites
and
CERCLA
NPL
sites.
Furthermore,
releases
have
been
reported
through
the
TRI.

Hexachlorobutadiene
has
also
been
detected
in
PWS
samples
collected
under
SDWA.
Occurrence
estimates
are
low
for
Round
1
and
Round
2
monitoring
with
only
0.13
%
and
0.05%
of
all
samples
showing
detections,
respectively.
Significantly,
the
values
for
the
99th
percentile
and
median
concentrations
of
all
samples
are
less
than
the
Minimum
Reporting
Level.
For
Round
1
samples
with
detections,
the
median
concentration
is
0.25
µ
g/
L
and
the
99th
percentile
concentration
is
10
µ
g/
L.
Median
and
99th
percentile
concentrations
for
Round
2
detections
are
0.30
µ
g/
L
and
1.5
µ
g/
L,
respectively.
Systems
with
detections
only
constitute
0.4%
of
Round
1
systems
and
0.2%
for
Round
2.
National
estimates
for
the
population
served
by
PWSs
with
detections
are
also
low,
especially
for
detections
greater
than
the
Health
Reference
Level.
For
both
rounds,
these
estimates
are
less
than
0.5%
of
the
national
population
(
Round
1:
781,076;
Round
2:
9,721).
HCBD
 
February
2003
5­
1
5.0
EXPOSURE
FROM
MEDIA
OTHER
THAN
WATER
This
section
describes
studies
which
measured
concentrations
of
HCBD
in
food,
air,
and
soil.
Exposure
of
adults
and
children
is
estimated
by
combining
the
reported
concentrations
with
the
estimated
intake
of
each
medium.
These
calculations
enable
a
comparison
of
exposure
to
HCBD
from
air,
food,
and
soil
with
that
anticipated
from
ingestion
of
drinking
water
(
see
Chapter
9.0).
Estimates
of
human
exposure
to
HCBD
via
food
and
air
have
previously
been
calculated
by
U.
S.
EPA
(
1998a).

5.1
Exposure
from
Food
Food
may
be
contaminated
with
HCBD
via
environmental
sources
or
by
contact
with
contaminated
water
during
food
processing
activity
(
DiNovi,
1997).
According
to
the
Food
and
Drug
Administration
(
FDA),
there
are
no
approved
uses
of
HCBD
either
directly
or
indirectly
in
foods,
including
food
processing
equipment
(
DiNovi,
1997).
HCBD
is
not
regulated
in
plastics.

5.1.1
Concentrations
in
Non­
Fish
Food
Items
Two
reports
provide
data
for
the
concentration
of
HCBD
in
food
items.
Yip
(
1976)
measured
HCBD
in
food
items
within
a
25­
mile
radius
of
tetrachloroethylene
and
trichloroethylene
manufacturing
plants
that
emit
HCBD
as
a
waste
product.
No
HCBD
was
detected
in
15
egg
samples
and
20
vegetable
samples.
One
of
20
milk
samples
contained
1.32
mg/
kg
HCBD.
Resampling
in
the
same
area
revealed
no
further
detections
in
milk,
raising
the
possibility
that
the
concentration
of
1.32
mg/
kg
measured
in
the
original
data
set
was
an
artifact.
This
study
reported
two
detection
limits
for
HCBD:
0.005
mg/
kg
for
nonfatty
foods
and
0.04
mg/
kg
for
fatty
foods.
Based
on
information
supplied
by
Kusznesof
(
1997),
U.
S.
EPA
(
1998a)
concluded
that
more
than
30%
of
foods
may
be
considered
fatty
foods
for
the
purpose
of
estimating
exposure
from
food
(
see
Section
5.1.3).

IARC
(
1979)
reported
concentrations
of
HCBD
in
foods
sampled
in
the
United
Kingdom.
HCBD
was
found
at
concentrations
of
0.00008
mg/
kg
in
fresh
milk,
0.002
mg/
kg
in
butter,
0.0002
mg/
kg
in
cooking
oil,
0.0002
mg/
kg
in
light
ale,
0.0008
mg/
kg
in
tomatoes,
and
0.0037
mg/
kg
in
black
grapes
(
IARC,
1979).

5.1.2
Concentrations
in
Fish
Concentrations
of
HCBD
in
fish
have
been
reported
in
multiple
studies.
Tchounwou
et
al.
(
1998)
demonstrated
that
aquatic
organisms,
particularly
fish,
may
be
a
significant
source
of
HCBD
transmission
from
contaminated
wetlands
to
humans.
Tissue
concentrations
of
HCBD
were
226.33
±
778.40
ng/
g
in
fish
collected
from
a
contaminated
study
site
in
Louisiana
and
6.84
±
10.41
ng/
g
in
fish
collected
from
the
corresponding
control
site.

In
other
studies,
fish
samples
from
the
Mississippi
River
were
reported
to
contain
HCBD
levels
ranging
from
100
to
4,700
ng/
g
(
Laska
et
al.,
1976;
Yip,
1976;
Yurawecz
et
al,
1976).
Levels
of
HCBD
generally
were
not
detected
in
fish
from
the
Great
Lakes
(
Camanzo
et
al.,
1987;
DeVault,
1985),
with
the
exception
of
trout
from
Lake
Ontario,
which
were
reported
to
contain
60
to
300
ng/
g
HCBD
 
February
2003
5­
2
(
Oliver
and
Nimi,
1983).
HCBD
was
not
detected
in
51
biota
samples
catalogued
in
the
STORET
database
(
Staples
et
al.,
1985).

The
National
Study
of
Chemical
Residues
in
Fish
(
NSCRF),
conducted
by
EPA's
Office
of
Water,
was
undertaken
to
determine
the
occurrence
of
selected
pollutants
in
fish
from
various
locations
across
the
United
States.
Pollutants
were
measured
in
bottom­
feeding
and
game
fish
at
nearly
400
sites
between
1986
and
1989
(
Kuehl
et
al.,
1994).
A
complete
presentation
of
the
study
plan
and
results
is
contained
in
a
joint
Office
of
Water
and
Office
of
Research
and
Development
report
(
U.
S.
EPA,
1992a).
To
obtain
nationwide
coverage,
samples
were
collected
at
sites
near
potential
point
and
nonpoint
pollution
sources,
at
background
sites
in
areas
generally
without
pollution
sources,
and
at
a
few
sites
from
the
U.
S.
Geological
Survey's
National
Stream
Quality
Accounting
Network
(
NASQAN).
Targeted
sites
were
chosen
near
areas
of
significant
industrial,
urban,
or
agricultural
activities,
including
more
than
100
sites
near
pulp
and
paper
mills.

Fish
species
chosen
for
sampling
were
those
routinely
consumed
by
humans
and/
or
those
expected
to
bioaccumulate
organic
contaminants.
At
most
locations,
the
NSCRF
analyzed
one
composite
sample
of
bottom­
feeding
fish,
usually
composed
of
whole­
body
samples.
Some
bottomfeeding
fish
composite
samples
were
composed
of
fillets.
In
areas
where
whole­
body
concentrations
were
high,
composite
samples
of
game
fish
were
usually
composed
of
fillets.
Each
composite
sample
contained
approximately
three
to
five
adult
fish
of
similar
size
from
the
site.
Pollutant
concentrations
were
measured
in
units
of
wet
weight
(
U.
S.
EPA,
1992a).

HCBD
was
detected
in
fish
at
3%
of
the
362
sites
sampled.
Fillet
samples
were
taken
from
106
sites.
The
mean
and
standard
deviation
of
HCBD
fish
concentrations
at
all
sites
were
0.6
ng/
g
and
8.7
ng/
g,
respectively
(
Kuehl
et
al.,
1994).
These
statistics
represent
the
overall
mean
from
all
samples,
not
just
from
the
positive
samples.
Concentrations
were
above
2.5
ng/
g
at
only
four
sites,
which
were
all
near
organic
chemical
manufacturing
plants
(
U.
S.
EPA,
1992a).
The
concentrations
observed
at
these
four
sites
are
provided
in
Table
5­
1.

The
methods
for
determining
the
mean
and
standard
deviation
for
HCBD
concentration
and
for
evaluating
samples
below
the
analytical
detection
limit
were
not
specifically
stated
by
U.
S.
EPA
(
1992a).
The
value
of
the
detection
limit
for
HCBD
was
not
given
in
U.
S.
EPA
(
1992a)
or
Kuehl
et
al.
(
1994).
However,
in
the
Kuehl
et
al.
(
1994)
study,
the
mean
concentration
was
calculated
using
one­
half
of
the
detection
limit
concentration
when
the
analyte
was
not
detected.
The
raw
data
for
HCBD
were
not
presented.

Hendricks
et
al.
(
1998)
evaluated
HCBD
levels
in
zebra
mussel
(
Dreissena
polymorpha)
and
eel
(
Anguilla
anguilla)
from
approximately
30
locations
in
the
Rhine­
Meuse
river
basin.
In
zebra
mussel,
HCBD
levels
were
240
ng/
kg
at
a
background
location
and
ranged
from
950
to
14,000
ng/
kg
wet
weight
within
the
study
area.
In
eel,
HCBD
levels
were
found
to
range
from
5,000
to
55,000
ng/
kg
wet
weight
within
the
study
area.
HCBD
 
February
2003
5­
3
Table
5­
1.
HCBD
Tissue
Concentration
in
Fish
Collected
Near
Four
Chemical
Manufacturing
Plants.

CONCENTRATION
(
ng/
g
wet
weight)
TYPE
OF
SAMPLE
LOCATION
164.0
Sea
Catfish
­
Whole
Body
Louisiana
23.0
Sea
Catfish
­
Whole
Body
Texas
10.50
Catfish
­
Fillet
Illinois
2.54
Catfish
­
Whole
Body
Louisiana
source:
U.
S.
EPA
(
1992a)

5.1.3
Intake
of
HCBD
from
Food
Non­
fish
Dietary
Intake
As
noted
above,
HCBD
has
been
found
in
a
variety
of
foods
in
the
United
Kingdom.
In
addition,
HCBD
may
have
been
incorrectly
measured
in
one
milk
sample
in
the
study
by
Yip
(
1976).
It
is
also
possible
that
HCBD
could
be
found
in
measurable
quantities
in
the
United
States.
However,
because
HCBD
was
generally
undetected
in
samples
taken
from
areas
within
25
miles
of
emission
sources,
U.
S.
EPA
(
1998a)
concluded
that
it
is
appropriate
to
assume
that,
on
average,
HCBD
will
not
be
found
in
food
at
detectable
levels.
Given
this
observation,
along
with
the
fact
that
HCBD
has
no
approved
uses
in
food,
it
is
anticipated
that
there
would
typically
be
no
chronic
exposure
to
HCBD
via
non­
fish
dietary
foods
(
U.
S.
EPA,
1998a).
Therefore,
the
average
estimate
of
HCBD
intake
from
non­
fish
foods
is
assumed
to
be
zero
(
U.
S.
EPA,
1998a).

A
high­
end
estimate
of
HCBD
exposure
may
be
made
by
assuming
a
concentration
of
onehalf
the
detection
limit
(
U.
S.
EPA,
1999b).
Because
the
percentages
of
fatty
or
non­
fatty
foods
in
the
diet
are
not
known
with
certainty,
a
conservative
estimate
is
made
using
one­
half
the
detection
limit
of
0.04
mg/
kg
noted
for
fatty
foods
in
Yip
(
1976).
The
resulting
concentration
of
0.02
mg/
kg
is
multiplied
by
an
estimate
of
total
daily
food
intake
of
2.6
kg/
day
and
divided
by
70
kg
to
obtain
a
total
daily
intake
of
HCBD
from
food
of
7.4
×
10­
4
mg/
kg­
day
in
adults.
For
children,
the
resulting
concentration
of
0.02
mg/
kg
was
multiplied
by
an
estimate
of
total
daily
food
intake
of
0.84
kg/
day
(
U.
S.
EPA,
1988)
and
divided
by
a
body
weight
of
10
kg
to
obtain
a
total
daily
intake
of
HCBD
from
food
of
1.68
×
10­
3
mg/
kg­
day.
For
the
majority
of
regions
of
the
United
States
in
which
HCBD
is
not
found,
using
one­
half
the
detection
limit
will
overestimate
the
amount
of
HCBD
in
food
(
U.
S.
EPA,
1999b).

Because
the
data
on
concentrations
in
food
are
limited,
and
because
the
implications
of
assuming
that
HCBD
occurs
at
one
half
the
detection
limit
for
fatty
foods
are
large,
further
research
may
be
required
to
refine
this
estimate.
HCBD
 
February
2003
5­
4
Fish
Dietary
Intake
U.
S.
EPA
(
1998a)
estimated
HCBD
intake
from
fish
using
the
tissue
concentration
data
from
Kuehl
et
al.
(
1994).
Because
these
data
were
taken
from
many
monitoring
stations
throughout
the
United
States,
the
estimate
may
be
reasonably
indicative
of
the
magnitude
of
intake
from
fish
consumption
when
HCBD
is
present
in
fish
tissue.
An
average
estimate
of
adult
exposure
was
obtained
by
multiplying
the
mean
concentration
of
0.6
ng/
g
from
the
Kuehl
et
al.
(
1994)
data
by
a
fish
intake
of
18
g/
day
for
the
general
population
and
dividing
by
a
body
weight
of
70
kg.
The
resulting
estimate
is
1.54
×
10­
7
mg/
kg­
day.
The
maximum
concentration
detected
in
fish
by
Kuehl
et
al.
(
1994)
can
also
be
used
to
estimate
the
high­
end
intake.
Following
the
same
procedure
above,
but
substituting
a
concentration
of
164
ng/
g,
one
obtains
a
high­
end
intake
of
4.22
×
10­
5
mg/
kg­
day.

An
average
estimate
of
HCBD
exposure
in
children
was
determined
by
multiplying
the
mean
concentration
of
0.6
ng/
g
from
the
Kuehl
et
al.
(
1994)
data
by
a
fish
intake
of
4
g/
day
for
the
general
population
and
dividing
by
a
body
weight
of
10
kg.
The
resulting
estimate
is
2.4
×
10­
7
mg/
kgday
The
maximum
concentration
detected
in
fish
by
Kuehl
et
al.
can
also
be
used
to
calculate
a
high­
end
estimate
of
intake
in
children.
Following
the
same
procedure
above,
but
substituting
a
concentration
of
164
ng/
g,
results
in
an
intake
of
4.37
×
10­
5
mg/
kg­
day.

5.2
Exposure
from
Air
5.2.1
Concentration
of
HCBD
in
Air
Concentration
data
for
HCBD
in
air
have
previously
been
summarized
by
U.
S.
EPA
(
1998a).
The
largest
compilation
of
data
on
ambient
air
concentrations
is
available
from
Shah
and
Heyerdahl
(
1988).
Shah
and
Heyerdahl
compiled
ambient
air
monitoring
data
for
volatile
organic
compounds
for
the
period
from
1970
to
1987.
A
total
of
72
observations
from
six
studies
were
reported
for
HCBD.
In
cases
where
more
than
one
sample
was
taken
per
day,
the
concentrations
were
averaged
and
weighted
by
sampling
time
when
the
sampling
time
varied
throughout
the
day.
When
more
than
one
sample
was
included
in
the
average,
values
less
than
the
minimum
quantifiable
limit
(
MQL)
were
included
as
one­
half
the
MQL
when
the
MQL
was
given.
When
the
MQL
was
not
indicated
in
the
Shah
and
Heyerdahl
study,
values
less
than
the
MQL
were
included
as
zeros
in
the
average.
If
the
resulting
average
was
less
than
the
MQL,
a
zero
was
included.
If
the
average
was
greater
than
the
MQL,
the
calculated
average
was
used.

As
reported
in
U.
S.
EPA
(
1998a),
the
average
and
median
of
all
ambient
HCBD
air
concentrations
measured
by
Shah
and
Heyerdahl
(
1988)
were
0.036
parts
per
billion
(
ppb)
(
0.42
:
g/
m3)
and
0.003
ppb
(
0.04
:
g/
m3),
respectively.
The
25th
and
75th
percentiles
were
0.001
ppb
(
0.01
:
g/
m3)
and
0.006
ppb
(
0.07
:
g/
m3).
Only
median
values
were
reported
for
urban
areas
and
sourcedominated
areas.
Of
56
samples
taken
from
urban
areas,
the
median
was
0.003
ppb
(
0.04
:
g/
m3).
Of
16
samples
taken
from
source­
dominated
areas,
the
median
was
0.002
ppb
(
0.02
:
g/
m3).
No
indoor
concentrations
were
reported
(
Shah
and
Heyerdahl,
1988).

Shah
and
Heyerdahl's
compilation
included
the
results
from
Pellizzari
et
al.
(
1979),
who
surveyed
the
occurrence
of
halogenated
hydrocarbons
in
various
environmental
media
of
five
HCBD
 
February
2003
5­
5
metropolitan
areas.
As
part
of
this
study,
HCBD
concentrations
in
the
vapor
phase
of
ambient
air
of
four
sites
were
compiled
from
other
research
programs,
as
well
as
from
monitoring
conducted
specifically
for
this
project.
In
the
Niagara
Falls
and
Buffalo,
New
York
area,
concentrations
were
found
to
range
from
trace
levels
to
389
ng/
m3,
with
six
of
15
determinations
(
40%)
containing
detectable
levels.
In
the
Baton
Rouge,
Louisiana
area,
two
of
11
determinations
(
18%)
were
18
and
37
ng/
m3.
Sampling
in
Houston,
Texas,
and
surrounding
areas
showed
a
range
of
trace
levels
to
2,066
ng/
m3,
with
seven
positive
values
from
a
total
of
17
determinations
(
41%).

Class
and
Ballschmiter
(
1987)
reported
that
the
troposphere
of
the
Northern
Hemisphere
contained
an
average
concentration
of
0.17
parts
per
trillion
(
ppt)
(
2
:
g/
m3)
HCBD
at
18
locations
sampled
from
1982
to
1986.
The
detection
limits
in
this
survey
were
between
0.01
and
0.1
ppt.

HCBD
concentrations
in
ambient
air
were
measured
in
two
studies
included
in
a
compilation
of
ambient
monitoring
data
for
the
Urban
Area
Source
Program
(
U.
S.
EPA,
1994).
In
the
first
survey,
concentrations
of
HCBD
were
reported
at
a
minimum
detection
level
of
540
:
g/
m3
when
measured
at
six
monitoring
stations
in
Columbus,
Ohio,
in
1989.
The
second
survey
was
conducted
in
Cincinnati,
Ohio,
from
1989
to
1991,
and
detected
HCBD
at
one
monitoring
site
at
a
concentration
of
1,000
:
g/
m3.

A
number
of
cities
had
HCBD
levels
ranging
from
2
to
11
ppt
(
0.02
to
0.12
:
g/
m3)
(
Pellizzari,
1978;
Singh
et
al.,
1980,
1982).
Niagara
Falls
had
higher
HCBD
levels,
with
concentrations
up
to
37
ppt
(
0.39
:
g/
m3)
found
in
ambient
air
levels
and
up
to
38
ppt
(
0.41
:
g/
m3)
found
in
the
basement
air
of
homes
near
industrial
and
chemical
waste
disposal
sites
(
Pellizzari,
1982).

However,
a
study
of
air
contaminants
in
Porto
Alegre,
Brazil
(
Grosjean
and
Rassmussen,
1999)
did
not
find
detectable
levels
of
HCBD
(
detection
limit
=
100
ppt)
at
any
of
46
sampling
locations.
A
monitoring
study
at
6
sampling
locations
in
Columbus,
Ohio
also
failed
to
detect
HCBD
in
the
air
(
Spicer
et
al.,
1996).

5.2.2
Intake
of
HCBD
from
Air
The
air
concentrations
reported
in
Shah
and
Heyerdahl
(
1988)
were
utilized
by
U.
S.
EPA
(
1998a)
to
calculate
an
estimate
of
exposure
because
this
data
set
included
a
reasonable
number
of
observations
(
n=
72).
For
adults,
the
mean
concentration
of
0.42
:
g/
m3
was
multiplied
by
an
average
air
intake
of
20
m3/
day
(
U.
S.
EPA,
1988).
The
resulting
value
was
divided
by
a
body
weight
of
70
kg,
and
the
units
were
converted
from
:
g
to
mg,
resulting
in
an
average
intake
of
1.2
×
10­
4
mg/
kgday
For
children,
the
mean
concentration
of
0.42
:
g/
m3
was
multiplied
by
an
average
air
intake
of
15
m3/
day
(
U.
S.
EPA,
1988).
The
resulting
value
was
divided
by
a
body
weight
of
10
kg
and
the
units
were
converted
from
:
g
to
mg,
resulting
in
an
intake
of
6.3
×
10­
4
mg/
kg­
day.
As
noted
in
U.
S.
EPA
(
1998a),
these
estimates
may
be
indicative
of
the
magnitude
of
HCBD
intake
from
air
in
urban
and
source
dominated
areas
where
the
chemical
is
present.
It
should
be
noted,
however,
that
these
concentration
data
are
older
than
data
from
the
Urban
Area
Source
Program
(
U.
S.
EPA,
1994)
and
Class
and
Ballschmiter
(
1987).
In
addition,
the
number
of
geographic
areas
sampled
throughout
the
United
States
by
Shah
and
Heyerdahl
was
not
indicated.
HCBD
 
February
2003
5­
6
5.3
Exposure
from
Soil
5.3.1
Concentration
of
HCBD
in
Soil
and
Sediment
Limited
data
on
the
concentrations
of
HCBD
in
soil
was
identified
in
a
RCRA
corrective
measures
study,
and
ranged
from
0.043
­
0.35
mg/
kg
(
USDOE,
2001)
.
Sediments
adsorb
HCBD
from
contaminated
water.
As
reported
in
U.
S.
EPA
(
1999b),
HCBD
was
not
detectable
in
any
of
the
196
sediment
samples
reported
in
the
STORET
database,
based
on
a
detection
limit
of
500
:
g/
kg
for
the
analyses
(
Staples
et
al.,
1985).
Sediments
from
the
Niagara
River
contained
2.9
to
11
:
g/
kg
HCBD
(
Oliver
and
Bourbonniere,
1985).
Sediments
from
the
Great
Lakes
contain
HCBD
typically
ranging
from
0.08
to
120
:
g/
kg
(
McConnell
et
al.,
1975).
Suspended
solids
collected
from
the
Rhine­
Meuse
river
basin
indicate
HCBD
levels
ranging
from
<
3.4
to
19
:
g/
kg
(
Hendriks
et
al.,
1998);
which
is
comparable
to
the
previously
mentioned
data
for
sediments
collected
in
the
United
States.

Several
studies
have
investigated
HCBD
levels
in
sediments
from
sites
in
Louisiana.
HCBD
levels
in
sediment
samples
from
a
Louisiana
swamp
environment
ranged
from
less
than
0.05
:
g/
kg
to
0.40
:
g/
kg
(
Abdelghani
et
al,.
1995).
These
concentrations
were
well
below
the
action
levels
of
4,000
:
g/
kg
for
sediment
(
U.
S.
EPA,
1991a).
At
a
Federal
Superfund
site
near
Baton
Rouge,
Louisiana,
preliminary
data
from
a
sampling
of
sediments
showed
HCBD
levels
from
2
to
3,770
mg/
kg
(
U.
S.
EPA,
1992b).
The
HCBD
level
in
a
sediment
sample
from
Lake
Charles,
Louisiana
was
found
to
be
3,500
:
g/
kg
(
Chen
et
al.,
1999).
A
sediment
sample
collected
from
the
intersection
of
an
industrial
canal
and
Bayou
d'Inde
(
a
tributary
of
the
Calcasieu
River
near
Lake
Charles)
and
analyzed
via
Soxhlet
extraction
was
found
to
contain
HCBD
at
a
level
of
17,200
±
1,000
:
g/
kg
(
Prytula
and
Pavlostathis,
1996).
Another
reported
sediment
sample
collected
from
this
industrial
canal
area
had
an
HCBD
level
of
36,000
±
6,900
:
g/
kg
(
Gess
and
Pavlostathis,
1997).
A
third
study
of
sediments
from
Bayou
d'Inde
found
levels
of
HCBD
ranging
from
1,550
to
8,220,000
:
g/
kg
of
organic
carbon.
Assuming
an
organic
carbon
content
in
the
sediment
of
1%,
this
level
is
equivalent
to
sediment
concentrations
of
15
to
82,200
:
g/
kg.

5.3.2
Intake
of
HCBD
from
Soil
Because
no
data
were
available
on
the
concentration
of
HCBD
in
soil
or
dust,
intake
from
soil
was
not
estimated.

5.4
Other
Residential
Exposures
HCBD
was
not
detected
in
sewage
influents
(
Levins
et
al.,
1979)
or
in
sewage
samples
(
U.
S.
EPA,
1990).
No
other
information
on
exposure
via
other
potentially
complete
residential
pathways
was
identified.

5.5
Summary
Estimated
mean
concentration
and
average
intake
values
for
HCBD
in
media
other
than
water
are
summarized
in
Table
5­
2.
Assuming
that
there
is
no
chronic
exposure
of
the
general
population
HCBD
 
February
2003
5­
7
Table
5­
2.
Summary
of
Concentration
Data
and
Exposure
Estimates
for
Media
Other
Than
Water.

PARAMETER
MEDIUM
Food
Air
Mean
Concentration
in
medium
Non­
fish
(
NF):
nondetect
Fish
(
F):
0.6
ng/
g
0.42
:
g/
m3
Estimated
average
daily
intake
(
mg/
kg­
day)
Adult
Child
Adult
Child
NF:
0
F:
1.5
×
10­
7
NF:
0
F:
2.4
×
10­
7
1.2
×
10­
4
6.3
×
10­
4
to
HCBD
from
non­
fish
dietary
sources,
inspection
of
the
data
indicates
that
most
intake
of
HCBD
by
the
general
population
occurs
via
inhalation.
However,
it
should
be
cautioned
that
this
preliminary
conclusion
is
subject
to
a
number
of
uncertainties:
1)
the
database
for
the
occurrence
of
HCBD
in
media
other
than
water
is
limited;
2)
many
of
these
data
are
more
than
20
years
old;
3)
in
some
cases,
information
on
the
geographic
location
of
sample
collection
or
analytical
details
are
lacking;
and
4)
data
for
HCBD
in
soil
and
dust
were
not
available
to
estimate
via
this
pathway.
HCBD
 
February
2003
6­
1
6.0
TOXICOKINETICS
This
section
describes
the
absorption,
distribution,
metabolism,
and
excretion
of
hexachlorobutadiene.
The
information
in
this
section
focuses
on
findings
in
animals
exposed
primarily
via
the
oral
route.
No
studies
were
identified
that
evaluated
the
toxicokinetic
behavior
of
HCBD
in
humans.
The
development
of
a
physiologically
based
pharmacokinetic
(
PBPK)
model
would
be
desirable
for
summarizing
and
extrapolating
the
toxicokinetic
behavior
of
HCBD
between
high­
dose
animal
studies
and
low­
dose
human
exposure.
In
order
to
develop
such
a
model,
blood
or
plasma
time­
course
data
for
HCBD
and,
if
possible,
its
metabolites,
would
be
required.
However,
no
pharmacokinetic
parameters
(
half­
life,
partition
coefficients,
etc.)
for
HCBD
or
its
metabolites
have
been
published
in
the
peer­
reviewed
literature,
and
no
blood
or
plasma
time­
course
for
HCBD
or
its
metabolites
has
been
published
for
any
species.
This
remains
as
an
important
gap
in
the
database
for
this
chemical.

6.1
Absorption
HCBD
is
readily
absorbed
following
oral
administration
to
experimental
animals.
Although
no
studies
have
quantitatively
determined
the
rate
of
absorption
of
HCBD
following
oral
dosing,
useful
information
has
been
obtained
from
studies
that
evaluated
the
distribution
and
excretion
of
this
compound.
Reichert
et
al.
(
1985)
administered
1
mg/
kg
of
14C­
HCBD
to
female
Wistar
rats
via
gavage.
The
compound
was
administered
in
a
tricaprylin
suspension
to
accommodate
its
low
water
solubility.
Approximately
76%
of
the
radioactivity
was
excreted
as
metabolites
in
the
urine,
feces,
or
expired
air
within
72
hours
after
administration,
suggesting
that
most
of
the
dose
was
absorbed.
When
a
higher
dose
of
50
mg/
kg
14C­
HCBD
was
administered
in
the
same
study,
69%
of
the
radioactivity
was
found
in
the
feces
and
was
predominantly
associated
with
unchanged
HCBD.
Just
11%
of
the
administered
radioactivity
was
excreted
in
the
urine
for
the
high­
dose
group,
compared
to
31%
for
the
low­
dose
animals.
The
study
authors
concluded
that
absorption
of
HCBD
was
saturated
in
animals
in
the
higher­
dose
group
(
Reichert
et
al.
1985;
U.
S.
EPA,
1991a).

Nash
et
al.
(
1984)
administered
200
mg/
kg
14C­
HCBD
via
oral
gavage
in
corn
oil
to
male
Wistar­
derived
rats.
Animals
were
sacrificed
2,
4,
8,
or
16
hours
after
dosing,
and
the
fate
of
the
administered
radioactivity
was
evaluated
using
whole­
body
autoradiographs.
The
investigators
reported
that
absorption
was
virtually
complete
within
16
hours
after
dosing.

Payan
et
al.
(
1991)
administered
1
mg/
kg
and
100
mg/
kg
14C­
HCBD
to
male
Sprague­
Dawley
rats,
using
an
aqueous
polyethylene
glycol
vehicle,
and
found
that
18.5
and
8.9%
of
the
administered
radioactivity,
respectively,
was
excreted
over
72
hours
in
the
urine.
Since
urinary
excretion
at
a
dose
of
1
mg/
kg
in
the
Reichert
et
al.
(
1985)
study
was
31%,
these
data
suggest
that
gastrointestinal
absorption
of
HCBD
was
greater
when
administered
in
a
lipophilic
vehicle
(
tricaprylin)
than
with
an
aqueous
vehicle
(
aqueous
polyethylene
glycol),
although
it
could
also
be
due
to
differences
in
animal
sex
and
strain
(
female
Wistar
rats
vs.
male
Sprague­
Dawley
rats).
As
noted
for
other
unsaturated
chlorinated
compounds,
HCBD
absorption
presumably
occurs
by
passive
diffusion
across
the
lipid
portion
of
the
intestinal
matrix
rather
than
by
active
or
protein­
facilitated
transport
(
ATSDR,
1994).
HCBD
 
February
2003
6­
2
Little
information
is
available
regarding
HCBD
absorption
following
exposure
by
other
routes.
Although
no
studies
were
located
that
described
absorption
in
humans
or
animals
after
inhalation
exposure
to
this
compound,
the
occurrence
of
systemic
effects
following
exposure
indicates
that
absorption
occurs
by
this
route
(
ATSDR,
1994).
With
regard
to
dermal
exposure,
Duprat
and
Gradiski
(
1978)
applied
doses
of
419
to
1,675
mg/
kg
HCBD
to
the
skin
of
rabbits
under
occluded
conditions,
and
reported
that
the
compound
was
completely
absorbed
within
8
hours.

6.2
Distribution
HCBD
has
been
detected
in
the
adipose
tissue
of
human
cadavers
at
concentrations
of
0.003
±
0.001
:
g/
g
(
Mes
et
al.,
1985),
but
not
whole
blood
in
any
of
36
residents
of
Love
Canal,
New
York
or
in
any
of
12
laboratory
volunteers
(
Bristol
et
al.,
1982).
Olea
et
al.
(
1999)
detected
HCBD
in
the
adipose
tissue
of
13
of
50
children
living
in
an
agricultural
region
of
southern
Spain.
The
mean
concentration
in
the
13
children
was
0.70
:
g
HCBD/
g
of
fat
(
range:
0.23
to
2.43
:
g
HCBD/
g
of
fat).
No
data
were
available
concerning
the
route
of
exposure.

Following
oral
administration,
HCBD­
related
radioactivity
preferentially
distributed
to
the
kidney,
liver,
adipose
tissue
and
brain
of
experimental
animals
(
Reichert,
1983;
Reichert
et
al.,
1985;
Dekant
et
al.,
1988a).
Covalent
binding
of
HCBD­
related
radioactivity
to
tissue
proteins
in
female
Wistar
rats
was
highest
during
the
first
six
hours
after
dosing,
and
was
higher
in
the
kidney
than
in
the
liver.
This
effect
was
independent
of
dose
(
Reichert
et
al.,
1985).
In
female
Wistar
rats
administered
1
mg/
kg
14C­
HCBD,
covalent
binding
of
the
radioactivity
to
protein
in
the
kidney
was
about
twice
that
in
the
liver
72
hours
after
dosing
(
Reichert
et
al.,
1985).
Nash
et
al.
(
1984)
reported
a
specific
localization
of
administered
radioactivity
in
male
Wistar­
derived
rats
in
the
outer
medulla
of
the
kidney,
as
revealed
by
autoradiographic
analysis
following
an
oral
dose
of
200
mg/
kg
14CHCBD
Payan
et
al.
(
1991)
conducted
a
study
in
male
Sprague­
Dawley
rats
in
which
the
bile
ducts
of
one
group
of
animals
administered
an
oral
dose
of
100
mg/
kg
14C­
HCBD
were
cannulated
so
that
bile
secretions
from
these
animals
could
be
infused
directly
into
the
duodenum
of
another
group
of
animals.
In
both
groups,
the
kidneys
contained
about
twice
as
much
radiolabel
as
the
liver.

No
studies
were
located
regarding
the
distribution
of
HCBD
in
humans
or
animals
after
inhalation
or
dermal
exposure.
Davis
et
al.
(
1980)
administered
0.1
mg/
kg
radiolabeled
14C­
HCBD
as
a
tracer
dose
to
a
control
group
of
male
Sprague­
Dawley
rats
(
5
animals/
group)
via
intraperitoneal
injection.
Another
group
received
the
same
amount
of
labeled
HCBD
plus
a
nephrotoxic
dose
of
300
mg/
kg
non­
labeled
HCBD.
The
highest
concentrations
of
radiolabel
were
found
in
the
liver,
kidney,
and
adipose
tissue
48
hours
after
administration.
Approximately
2.6
and
2.3%
of
the
administered
14C
radiolabel
were
retained
in
the
livers
of
low­
and
high­
dose
animals,
respectively.
The
fraction
of
the
tracer
retained
in
kidney
varied
from
2.5%
at
the
low
dose
to
0.5%
at
the
high
dose.
The
fraction
of
the
dose
found
in
adipose
tissue
was
not
determined.
Very
low
levels
of
the
radiolabel
(
less
than
0.2%)
were
found
in
the
brain,
lung,
heart,
and
muscle.
HCBD
 
February
2003
6­
3
6.3
Metabolism
No
available
studies
have
characterized
the
metabolism
of
HCBD
in
animals
following
inhalation
or
dermal
exposure.
The
metabolism
of
HCBD
in
animals
has
been
studied
in
isolated
hepatocytes
from
male
Sprague­
Dawley
rats
(
Jones
et
al.,
1985)
and
by
characterization
of
metabolites
identified
in
urine,
bile,
and
feces
following
oral
exposure
to
the
compound
(
Figure
6­
1).
Following
ingestion
and
absorption
from
the
gastrointestinal
tract
of
mice
and
rats,
HCBD
is
initially
transported
to
the
liver,
where
it
is
conjugated
with
glutathione
to
form
S­(
1,1,2,3,4
­
pentachlorobutadienyl)
glutathione
in
a
reaction
mediated
by
glutathione
S­
transferase
(
Wolf
et
al.,
1984;
Garle
and
Fry,
1989;
Dekant
et
al.,
1988b;
Koob
and
Dekant,
1992).
No
studies
have
found
non­
enzymatic
glutathione­
conjugation
of
HCBD,
and
none
have
investigated
GST
subtype
specificity
for
HCBD.
In
male
Sprague­
Dawley
rats,
a
di­
substituted
glutathione
conjugate,
1,4
bis(
1,2,3,4­
tetrachlorobutadienyl)
glutathione,
is
also
formed
in
the
liver
(
Jones
et
al.,
1985),
whereas
in
mice,
only
the
mono­
substituted
conjugate
is
produced
(
Dekant
et
al.,
1988a).
The
glutathione
conjugate
is
then
excreted
in
bile
and
transported
back
into
the
gastrointestinal
tract
(
Koob
and
Dekant,
1992).
Nash
et
al.
(
1984),
for
example,
collected
bile
excretions
from
cannulated
male
Wistar­
derived
rats
that
had
been
orally
administered
200
mg/
kg
14C­
HCBD
and
determined
that
40%
of
the
bile
radioactivity
was
associated
with
the
glutathione
conjugate.
Studies
conducted
by
Lock
et
al.
(
1984)
demonstrate
that
induction
or
inhibition
of
cytochrome
P450
metabolism
does
not
alter
the
nephrotoxicity
of
HCBD
in
male
or
female
mice
administered
24,
48,
96,
or
144
mol/
kg
HCBD
by
gavage.
Glutathione
and
N­
acetylcysteine
conjugates
of
HCBD
exhibit
more
potent
nephrotoxicity
in
male
and
female
mice
(
Lock
et
al.,
1984;
Ishmael
and
Lock,
1986)
and
male
Sprague­
Dawley
rats
(
Nash
et
al.,
1984)
than
unconjugated
HCBD.
This
suggests
that
conjugated
metabolites,
unlike
oxidative
metabolites,
may
play
a
large
role
in
HCBD
nephrotoxicity.

The
glutathione
conjugate
of
HCBD
can
be
reabsorbed
intact
from
the
gastrointestinal
tract
of
male
rats
(
Koob
and
Dekant,
1992;
Gietl
et
al.,
1991).
Alternatively,
a
portion
of
it
can
be
catabolized
by
(­
glutamyltranspeptidase
and
dipeptidases
in
the
rat
gastrointestinal
tract
to
the
cysteine
conjugate,
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine
(
Jones
et
al.,
1985;
Gietl
et
al.,
1991;
Koob
and
Dekant,
1992).
In
the
rat,
both
the
glutathione
and
cysteine
conjugates
are
subject
to
several
alternative
fates.
These
conjugates
may
be
reabsorbed
from
the
gut
and
be
translocated
to
the
kidney
(
Koob
and
Dekant,
1992),
undergo
enterohepatic
circulation
(
Nash
et
al.,
1984;
Gietl
et
al.,
1991;
Gietl
and
Anders,
1991),
or
be
excreted
with
the
feces
(
Dekant
et
al.
1988a).
However,
the
majority
of
the
glutathione
conjugate
is
delivered
to
the
kidney
by
systemic
circulation
in
the
rat
(
Koob
and
Dekant,
1992).
Working
with
isolated
perfused
rat
livers,
Koob
and
Dekant
(
1992)
determined
that
a
maximum
of
39%
of
the
glutathione
conjugate
was
recirculated
to
the
liver
of
rats,
whereas
up
to
79%
of
the
cysteine
conjugate
was
recirculated.
Nash
et
al.
(
1984)
reported
that
the
cysteine
and
glutathione
conjugates
represented
12%
and
40%,
respectively,
of
the
radioactivity
excreted
in
the
bile
of
cannulated
rats
orally
administered
200
mg/
kg
HCBD.
When
the
cysteine
conjugate
is
recirculated
to
the
rat
liver,
a
minor
fraction
of
this
metabolite
is
converted
by
Nacetyltransferase
to
an
acetylated
cysteine
conjugate,
N­
acetyl­
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
Lcysteine
(
N­
AcPCBC)
(
Koob
and
Dekant,
1992).
HCBD
 
February
2003
6­
4
NHCOCH
3
NHCOCH3
O
Cl
Cl
Cl
C
C
C
C
Cl
Cl
Cl
Cl
Cl
Cl
C
C
C
C
Cl
Cl
Cl
Cl
Cl
C
C
C
C
Cl
Cl
Cl
Cl
C
C
C
C
Cl
SG
Cl
Cl
Cl
Cl
C
C
C
C
Cl
SOH
Cl
SG
S
Cys
Gly
GS
Cl
Cl
C
C
C
C
Cl
Cl
Cl
Cl
C
C
C
C
Cl
Cl
Cl
Cl
Cl
C
C
C
S­
CH2­
CH
Cl
Cl
Cl
Cl
Cl
C
C
C
C
S­
CH2­
COOH
Cl
Cl
Cl
Cl
Cl
C
C
C
C
SH
Cl
Cl
Cl
Cl
Cl
C
C
C
C
Cl
S­
CH2­
CH
COOH
COOH
S
Hexachloro­
1,3­
butadiene
Glutathione
Transferase
GSH
Liver
S­(
1,1,2,3,4­
pentachloro
butadienyl)
glutathione
(
PCBG)

1,4
bis(
1,2,3,4­
tetrachloro
butadienyl)
glutathione
(
b)
­
glutamic
acid
1,1,2,3,4­
pentachloro
butadienyl
sulfenic
acid
N­
acetyl­
transferase
N­
acetyl­
S­(
1,1,2,3,4­

pentachlorobutadienyl
L­
cysteine
(
N­
AcPCBC)
­
lyase
deamination
decarboxylation
S­
methylation
1,1,2,3,4­
pentachlorobutadienyl
methylthioether
Thioketene
S­(
1,1,2,3,4­
pentachlorobutadienyl)­

L­
cysteine
(
PCBC)
1,1,2,3,4­
pentachloro
butadienylthiol
HCl
1,1,2,3,4­
pentachlorobutadienyl
carboxymethylthioether
­
glycine
Glutathione
Transferase
GSH
Liver
dipeptidase
Notes:

(
a)
Di­
substituted
compounds
are
formed
in
rats,
but
not
in
mice
(
b)
Metabolism
parallels
that
for
the
monosubstituted
compound
*
Adapted
from:
Dekant
et
al.,
1991;
Jaffe
et
al.,
1983;
Nash
et
al.,
1984;
Wolf
et
al.,
1984;
Jones
et
al.,
1985;
Reichert
et
al.,
1985;

Reichert
and
Schutz,
1986;
Wild
et
al.,
1986;
Birner
et
al.,
1998;
ATSDR,
1994.
deacetylase
Cyt.
P450
3A
Male­
specific
Sulfoxidation
pathway
 
­
glutamyl
transpeptidase
N­
acetyl­
S­(
1,1,2,3,4­
pentachlorobutadienyl)

­
L­
cysteine
sulfoxide
(
N­
AcPCBC­
SO)

covalent
binding
Cl
Cl
C
C
C
C
Cl
Cl
S­
CH2­
CH
Cl
toxicity
(
a)

(
a)
S­
CH
3
COOH
C
Cl
Cl
NH2
decarboxylation
 
GSH
=
glutathione
gly
=
glycine
cys
=
cysteine
Figure
6­
1.
Proposed
Pathways
for
Hexachlorobutadiene
Metabolism.
HCBD
 
February
2003
6­
5
Further
processing
of
both
the
cysteine
and
glutathione
conjugates
occurs
in
the
kidney,
which
possesses
high
(­
glutamyltranspeptidase
activities
in
the
brush­
border
membrane
of
the
proximal
tubular
cells
(
Dekant
and
Vamvakas,
1993;
Dekant
et
al.,
1990).
Renal
deacetylase,
(­
glutamyltranspeptidase,
and
dipeptidase
enzymes
convert
the
acetylated
cysteine
conjugate
and
the
glutathione
conjugate
to
the
cysteine
conjugate,
which
accumulates
in
the
kidney
(
Dekant
et
al.,
1990).
The
cysteine
conjugate
is
subsequently
activated
to
a
reactive
and
electrophilic
thioketene
intermediate
(
Dekant
et
al.,
1990;
Green
and
Odum
1985).
This
conversion
is
catalyzed
by
the
enzyme­
cysteine
conjugate
$­
lyase,
which
is
localized
in
the
cytosol
and
mitochondria
of
the
epithelial
cells
of
the
proximal
tubule
(
Lash
et
al.,
1986;
Stevens,
1985;
Stevens
et
al.
1986;
Jones
et
al.,
1988;
MacFarlane
et
al.,
1989;
Kim
et
al.,
1997).

Another
pathway
for
metabolic
disposition
of
the
cysteine
conjugate
of
HCBD
in
the
kidney
is
the
conversion
of
the
cysteine
conjugate
to
a
mercapturic
acid,
N­
acetyl­
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine,
by
the
renal
enzyme
N­
acetyltransferase
(
Birner
et
al.,
1997).
This
metabolite
is
excreted
in
the
urine,
accounting
for
10%
of
urinary
radioactivity
in
rats
orally
administered
100
mg/
kg
14C­
HCBD
(
Reichert
and
Schutz,
1986).
Other
pathways
that
result
in
the
excretion
of
the
cysteine
conjugate
involve
the
deamination
and
subsequent
decarboxylation
of
the
cysteine
conjugate,
resulting
in
the
formation
of
methylthiolated
metabolites
such
as
1,1,2,3,4­
pentachlorobutadiene
methylthioether
and
1,1,2,3,4­
pentachlorobutadiene
carboxymethylthioether
(
Reichert
et
al.,
1985).
1
to
8%
of
the
administered
radioactivity
is
recovered
as
carbon
dioxide
in
exhaled
air
from
rats
(
Reichert
et
al.,
1985;
Payan
et
al.,
1991),
probably
from
decarboxylation
of
the
cysteine
conjugate.

Evidence
for
a
male­
specific
HCBD
metabolic
pathway
in
rats
has
been
reported
by
Birner
and
colleagues
(
Birner
et
al.,
1995,
1998;
Werner
et
al.
1995a).
The
metabolite
N­
acetyl­
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine
sulfoxide
(
N­
AcPCBC­
SO)
is
detected
in
the
urine
of
male,
but
not
female,
rats
following
oral
administration
of
HCBD.
Formation
of
this
metabolite
is
mediated
by
cytochrome
P450
3A
monooxygenases,
which
are
expressed
only
in
male
rats
(
Birner
et
al.,
1995;
Werner
et
al.,
1995a).
This
metabolite
has
been
found
to
be
cytotoxic
to
proximal
tubular
cells
in
vitro
without
activation
by
$­
lyase
(
Birner
et
al.,
1995).
When
given
intravenously,
N­
AcPCBC­
SO
produced
necroses
of
the
kidney
tubules
in
male
rats
(
Birner
et
al.,
1998).

The
N­
AcPCBC­
SO
formed
in
male
rats
occurs
as
two
diastereomers
present
in
equimolar
amounts:
(
R)­
N­
AcPCBC­
SO
and
(
S)­
N­
AcPCBC­
SO
(
Werner
et
al.,
1995b).
These
compounds
are
structurally
analogous
to
unsaturated
carbonyl
compounds
and
thus
may
be
candidates
for
detoxification
via
glutathione
conjugation
(
Rosner
et
al.,
1998).
Experimental
evidence
obtained
in
vitro
suggests
that
glutathione
conjugation
of
the
two
diastereomers
is
catalyzed
by
different
glutathione
S­
transferases,
resulting
in
the
formation
of
different
products
(
Rosner
et
al.,
1998).
Incubation
of
the
(
R)­
sulfoxide
diastereomer
with
rat
liver
cytosol
resulted
in
formation
of
(
R)­
Nacetyl
S­(
4­
glutathion­
S­
yl­
1,2,3,4­
tetrachlorobutadienyl)­
L­
cysteine
sulfoxide.
Incubation
of
the
(
S)­
sulfoxide
produced
two
glutathione
conjugates
identified
as
(
S)­
N­
acetyl­
S­(
4­
glutathion­
S­
yl­
1,2,3,4­
tetrachlorobutadienyl)­
L­
cysteine
sulfoxide
and
(
S)­
N­
acetyl­
S­(
2­
glutathion­
S­
yl­
1,3,4,4­
tetrachlorobutadienyl)­
L­
cysteine
sulfoxide.
In
the
presence
of
rat
kidney
cytosol,
only
the
(
S)­
Nacetyl
S­(
2­
glutathion­
S­
yl­
1,3,4,4­
tetrachlorobutadienyl)­
L­
cysteine
sulfoxide
conjugate
was
formed.
Glutathione
conjugation
of
the
(
R)­
sulfoxide
was
not
observed.
The
observed
pattern
of
HCBD
 
February
2003
6­
6
product
formation
was
attributed
to
catalysis
by
different
glutathione
S­
transferases
in
liver
("­
and
:­
class)
and
kidney
("­
class).
This
hypothesis
was
confirmed
by
product
analysis
following
incubation
of
N­
AcPCBC­
SO
with
purified
rat
"­
and
:­
class
glutathione
S­
transferases.

Very
little
information
is
available
on
the
toxicokinetic
behavior
of
HCBD
in
humans.
However,
the
key
steps
in
the
metabolism
of
HCBD
have
been
examined
in
vitro
using
human
tissues.
The
human
liver
microsomal
glutathione
transferase
responsible
for
HCBD
conjugation
has
been
isolated
and
purified
(
McLellan
et
al.,
1989),
and
the
microsomal
enzyme
activity
is
40­
fold
higher
than
the
activity
detected
in
the
cytosol
(
Oesch
and
Wolf,
1989).
The
rate
of
enzymatic
formation
of
S­(
1,2,3,4,4­
pentachlorobutadienyl)
glutathione
(
PCBG)
from
HCBD
in
human
liver
cytosol
is
in
the
same
order
of
magnitude
as
the
rates
observed
in
rat
and
mouse
cytosol
(
Dekant
et
al.,
1998).

The
enzyme
(­
glutamyl
transpeptidase,
which
catalyzes
the
conversion
of
glutathione
Sconjugates
to
the
corresponding
cysteine
conjugates,
has
been
detected
in
human
tissues
(
Shaw
et
al.,
1978).
The
kidney­
to­
liver
activity
ratio
for
(­
glutamyl
transpeptidase
in
human
tissues
is
approximately
22,
which
is
comparable
to
the
ratios
observed
in
pig
and
guinea
pig.
However,
this
ratio
is
much
lower
than
in
rat,
where
a
kidney­
to­
liver
ratio
of
875
has
been
observed
(
Hinchman
and
Ballatori,
1990).

Cysteine
conjugate
$­
lyase
has
been
isolated
and
purified
from
human
kidney
cytosol
(
Lash
et
al.,
1990),
and
the
human
$­
lyase
gene
has
been
cloned
and
expressed
(
Perry
et
al.,
1995).
$­
lyase
activity
has
been
demonstrated
in
the
human
kidney
(
Green
et
al.,
1990)
and
human
proximal
tubular
cells
(
Chen
et
al.,
1990).
Collectively,
these
human
studies
suggest
that
humans
have
the
ability
to
metabolize
HCBD
to
toxic
metabolites.
However,
the
activity
of
HCBD
metabolizing
enzymes,
particularly
renal
$­
lyase,
may
be
many­
fold
lower
in
humans
than
the
corresponding
enzymes
in
rat
(
Lock,
1994;
Lash
et
al.,
1990;
Anders
and
Dekant,
1998).

Werner
et
al.
(
1995b)
demonstrated
that
human
liver
microsomes
are
capable
of
oxidizing
N­
acetyl­
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine
to
the
corresponding
sulfoxide
(
N­
AcPCBCSO
In
contrast
to
the
male­
specific
formation
of
N­
AcPCBC­
SO
in
rats
described
above,
formation
of
the
sulfoxide
was
detected
in
human
microsomes
prepared
from
both
male
and
female
donors.
Inhibitor
studies
suggest
that
formation
of
the
sulfoxide
is
catalyzed
by
members
of
the
cytochrome
P450
3A
subfamily.
Since
this
subfamily
constitutes
a
major
fraction
of
cytochrome
P450
content
in
human
liver,
the
formation
of
the
sulfoxide
is
expected
to
occur
in
humans
exposed
to
HCBD.
Incubation
of
N­
AcPCBC­
SO
with
purified
human
glutathione
S­
transferase
M1­
1
(:­
class)
catalyzes
the
formation
of
(
S)­
N­
acetyl­
S­(
4­
glutathion­
S­
yl­
1,2,3,4­
tetrachlorobutadienyl)­
Lcysteine
sulfoxide
and
(
R)­
N­
acetyl­
S­(
4­
glutathion­
S­
yl­
1,2,3,4­
tetrachlorobutadienyl)­
L­
cysteine
sulfoxide,
the
same
products
formed
in
the
presence
of
rat
:­
class
glutathione
S­
transferase
(
Rosner
et
al.,
1998).
HCBD
 
February
2003
6­
7
6.4
Excretion
HCBD
and
its
metabolites
are
excreted
in
urine,
feces,
and
exhaled
air.
In
rodents,
oral
exposure
to
HCBD
results
in
fairly
rapid
urinary
and
fecal
excretion.
The
half­
life
of
elimination
of
HCBD­
related
radioactivity
bound
to
tissue
proteins
was
22
hours
in
both
liver
and
kidney
(
Reichert
et
al.,
1985).
Total
excretion
within
72
hours
was
found
to
be
at
least
65%
of
a
single
oral
dose
of
up
to
100
mg/
kg
in
rats
and
mice
(
Reichert
and
Schutz,
1986;
Dekant
et
al.,
1988a).
At
higher
doses
(
30
to
200
mg/
kg),
5
to
11%
of
the
radiolabel
was
excreted
in
the
urine
(
Dekant
et
al.,
1988a;
Nash
et
al.,
1984;
Reichert
and
Schutz,
1986;
Reichert
et
al.,
1985;
Payan
et
al.,
1991),
while
a
dose
of
1
mg/
kg
resulted
in
urinary
excretion
of
approximately
18.5
to
30%
of
the
administered
radioactivity
(
Payan
et
al.,
1991;
Reichert
et
al.,
1985).
Payan
et
al.
(
1991)
attributed
the
fractional
decrease
in
urinary
excretion
with
increase
in
dosage
to
a
saturation
of
hepatobiliary
excretion
or
a
reduction
of
biliary
metabolite
reabsorption.
Exhaled
unchanged
labeled
HCBD
accounted
for
approximately
5%
of
oral
doses
of
1
and
50
mg/
kg
when
measured
within
72
hours
of
administration,
while
exhaled
labeled
carbon
dioxide
accounted
for
1%
or
4%
of
the
administered
dose
(
Reichert
et
al.,
1985).
The
authors
did
not
indicate
the
metabolic
pathway
for
carbon
dioxide
formation.

Fecal
excretion
is
the
main
pathway
of
elimination
for
HCBD
and
HCBD
metabolites.
Reichert
et
al.
(
1985)
reported
that
elimination
of
HCBD
in
the
feces
represented
42
or
69%
of
the
radioactivity
orally
administered
to
female
Wistar
rats
at
doses
of
1
or
50
mg/
kg,
respectively.
The
difference
in
recovery
was
attributed
to
an
apparent
absorption
saturation
threshold.
Dekant
et
al.
(
1988a)
found
that
67
to
77%
of
the
radioactivity
in
an
oral
30
mg/
kg
dose
in
corn
oil
was
present
in
the
feces
of
mice
72
hours
after
administration.
Payan
et
al.
(
1991)
administered
oral
doses
of
14CHCBD
in
polyethylene
glycol
to
male
rats.
After
72
hours,
the
feces
and
contents
of
the
gastrointestinal
tract
contained
62%
of
a
1
mg/
kg
dose
and
72%
of
a
100
mg/
kg
dose.

Enterohepatic
circulation
has
been
demonstrated
in
animals
following
oral
administration
of
HCBD.
Nash
et
al.
(
1984)
administered
200
mg/
kg
14C­
HCBD
to
rats
with
and
without
cannulated
bile
ducts.
Feces
and
urine
were
collected
over
a
5­
day
period.
Over
the
course
of
the
experiment,
the
non­
cannulated
animals
excreted
39%
of
the
administered
radioactivity
in
feces.
On
each
of
the
first
two
days
post­
dosing,
approximately
5%
and
3.5%
of
the
administered
radioactivity
were
found
in
the
feces
and
urine,
respectively.
In
contrast,
bile
excretions
collected
from
the
cannulated
animals
on
each
of
the
first
2
days
post­
dosing
contained
17
 
20%
of
the
administered
radioactivity
(
with
35%
total
excretion
by
this
route
over
two
days).
These
findings
indicate
extensive
reabsorption
of
biliary
metabolites.

Payan
et
al.
(
1991)
also
compared
excretion
patterns
in
bile­
duct
cannulated
and
noncannulated
rats
orally
dosed
with
1
mg/
kg
14C­
HCBD.
Urinary
excretion
after
72
hours
accounted
for
18%
of
the
administered
radioactivity
in
intact
animals,
but
just
11%
of
the
radioactivity
in
the
cannulated
rats.
In
comparison,
fecal
excretion
represented
62%
and
3%,
of
the
dose
administered
to
non­
cannulated
and
cannulated
animals,
respectively.
In
cannulated
rats,
67%
of
the
dose
was
excreted
into
the
bile.
When
bile
excretions
(
isolated
from
bile
duct­
cannulated
rats
orally
dosed
with
100
mg/
kg
HCBD)
were
directly
infused
into
the
duodenum,
approximately
80%
of
the
biliary
metabolites
are
reabsorbed,
with
only
20%
remaining
in
the
feces
and
gastrointestinal
tract.
HCBD
 
February
2003
6­
8
Several
studies
have
reported
the
identity
of
excreted
metabolites
following
exposure
to
14CHCBD
Metabolites
identified
in
the
urine
of
treated
rats
or
mice
include
S­(
1,1,2,3,4­
pentachlorobutadienyl)
glutathione,
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine,
N­
acetyl­
S­
(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine,
1,1,2,3,4­
pentachlorobutadienyl
sulfenic
acid,
1,1,2,3,4­
pentachlorobutadiene
methylthioether,
1,1,2,3,4­
pentachlorobutadiene
carboxymethylthioether,
and
1,1,2,3­
tetrachlorobutenoic
acid
(
Dekant
et
al.,
1988a;
Nash
et
al.,
1984;
Reichert
and
Schultz,
1986;
Reichert
et
al.,
1985).
As
noted
previously,
the
novel
metabolite
N­
acetyl­
S­(
1,1,2,3,4,­
pentachlorobutadienyl)­
L­
cysteine
sulfoxide
has
been
detected
in
the
urine
of
male,
but
not
female,
rats
following
oral
administration
of
HCBD
(
Birner
et
al.,
1995).

Comparatively
few
data
are
available
on
the
identity
of
fecal
metabolites.
Dekant
et
al.
(
1988a)
administered
a
single
30
mg/
kg
gavage
dose
of
14C­
HCBD
in
corn
oil
to
male
and
female
NMRI
mice.
The
feces
were
collected
over
a
72­
hour
period
following
dose
administration.
Approximately
80%
of
the
fecal
radioactivity
was
associated
with
HCBD.
About
10%
of
the
radiolabel
was
associated
with
the
HCBD
metabolite
S­(
1,1,2,3,4­
pentachlorobutadienyl)
glutathione.
The
remainder
of
the
fecal
radioactivity
was
present
as
polar
metabolites
which
could
not
be
structurally
identified.
HCBD
 
February
2003
7­
1
7.0
HAZARD
IDENTIFICATION
7.1
Human
Effects
Limited
information
is
available
on
the
human
health
effects
associated
with
exposure
to
HCBD.
A
review
of
the
available
literature
did
not
identify
case
reports
describing
the
outcome
of
accidental
or
intentional
HCBD
exposure,
or
reports
of
systemic
toxicity
following
oral
or
dermal
HCBD
exposure.
A
number
of
studies
have
evaluated
health
effects
in
workers
occupationally
exposed
to
HCBD
via
inhalation,
and
these
studies
are
described
below.

7.1.1
Short­
Term
Studies
No
short­
term
studies
describing
HCBD
health
effects
in
humans
were
located.

7.1.2
Long­
Term
and
Epidemiological
Studies
General
Population
No
general
population
studies
of
HCBD
toxicity
were
located.

Sensitive
Populations
No
studies
concerning
HCBD
toxicity
in
sensitive
populations
were
located.

Occupational
Exposure
Studies
German
(
1986)
conducted
two
cytogenetic
studies
of
workers
employed
in
an
HCBD
production
facility.
The
exposure
levels
reported
by
the
manufacturer
ranged
from
1.6
to
16.9
mg/
m3.
The
investigators
found
an
increased
frequency
of
chromosomal
aberrations
in
the
peripheral
lymphocytes
of
exposed
workers.
However,
the
frequency
of
aberrations
was
not
associated
with
duration
of
employment
in
the
HCBD
manufacturing
facility
(
WHO,
1994),
suggesting
that
factors
other
than
HCBD
exposure
contributed
to
the
observed
effects.

Additional
occupational
studies
have
evaluated
health
effects
in
workers
exposed
to
HCBD.
However,
in
each
case
concurrent
exposure
of
workers
to
other
chemicals
limits
the
usefulness
of
the
data
for
evaluation
of
HCBD
human
health
effects.
Krasniuk
et
al.
(
1969),
for
example,
evaluated
health
effects
in
153
farm
workers
intermittently
exposed
over
a
period
of
four
years
to
soil
and
grape
fumigants
containing
HCBD.
When
compared
to
a
control
population
of
52
unexposed
workers,
HCBD­
exposed
workers
exhibited
increased
incidence
of
arterial
hypotension,
myocardial
dystrophy,
chest
pains,
upper
respiratory
tract
changes,
liver
effects,
sleep
disorders,
hand
trembling,
nausea,
and
disordered
olfactory
functions
(
U.
S.
EPA,
1991a).
Interpretation
of
these
data
is
confounded
by
concurrent
exposure
of
these
workers
to
polychlorobutane­
80.
HCBD
 
February
2003
7­
2
Burkatskaya
et
al.
(
1982)
reported
adverse
health
effects
in
vineyard
workers
exposed
to
fumigants
containing
HCBD.
However,
the
role
of
HCBD
could
not
be
evaluated
because
the
workers
were
concurrently
exposed
to
other
agricultural
chemicals
(
WHO,
1994).

Driscoll
et
al.
(
1992)
determined
the
concentrations
of
individual
serum
or
plasma
bile
acids
in
workers
exposed
to
chlorinated
hydrocarbons,
including
HCBD,
carbon
tetrachloride,
and
perchloroethylene.
These
investigators
reported
increases
in
four
serum
bile
acid
parameters
in
workers
exposed
via
inhalation
to
0.005
 
0.02
ppm
HCBD.
The
study
found
no
significant
relation
between
bile
acid
parameters
or
liver
function
tests
and
exposure.
As
in
the
studies
above,
the
specific
contribution
of
HCBD
exposure
to
the
observed
effects
could
not
be
evaluated.

7.2
Animal
Studies
7.2.1
Acute
Toxicity
Oral
Exposure
Schwetz
et
al.
(
1977)
reported
a
single­
dose
oral
LD
50
value
(
the
dose
that
produces
lethality
in
50%
of
the
experimental
animals)
of
65
mg
HCBD/
kg
for
male
weanling
rats
and
46
mg/
kg
for
female
weanling
rats.
A
single­
dose
oral
LD
50
value
of
90
mg/
kg
was
reported
for
adult
rats
(
Kennedy
and
Graepel,
1991).
These
data
suggest
that
age
and
gender
may
be
significant
variables
in
the
acute
toxicity
of
HCBD.
Single­
dose
LD
50
values
reported
for
other
rodents
were
80
to
116
mg/
kg
for
mice
and
90
mg/
kg
for
guinea
pigs
(
U.
S.
EPA,
1991a).

Three
studies
have
evaluated
the
non­
lethal
acute
effects
of
oral
HCBD
exposure.
Nash
et
al.
(
1984)
administered
a
single
oral
dose
of
200
mg/
kg
HCBD
in
polyethylene
glycol
to
six
male
Wistar­
derived
rats.
Treatment
with
HCBD
increased
plasma
urea
concentration
and
decreased
plasma
alanine
aminotransferase
activity.
Analysis
of
urine
revealed
significantly
(
p
>
0.01)
increased
levels
of
glucose,
protein,
alkaline
phosphatase,
N­
acetyl­
ß­
D­
glucosaminidase
(
NAG),
(­
glutamyl
transpeptidase
and
alanine
aminopeptidase
over
control
levels.
These
are
commonly
used
biochemical
markers
of
kidney
damage.

Jonker
et
al.
(
1993a)
investigated
the
acute
oral
toxicity
of
HCBD
in
12­
week­
old
male
Wistar
rats.
The
investigators
administered
single
doses
of
0,
10,
100,
or
200
mg/
kg
HCBD
in
corn
oil
by
gavage
to
five
rats
per
treatment
group.
Urine
was
collected
at
intervals
of
0
to
6
and
6
to
24
hours.
All
rats
were
sacrificed
at
24
hours.
No
treatment­
related
effects
were
observed
at
the
10
mg/
kg
dose.
HCBD
induced
a
variety
of
adverse
effects
at
the
two
highest
dose
levels.
Kidney
weight,
blood
plasma
creatinine
level,
urinary
pH
and
occult
blood,
number
of
epithelial
cells
in
the
urine,
urinary
lactate
dehydrogenase
and
NAG
activity
were
significantly
increased
(
p<
0.05)
at
100
and
200
mg/
kg.
Additional
effects
observed
in
the
200
mg/
kg
dose
group
included
reduced
body
weight,
reduced
food
intake,
elevated
plasma
urea
level,
and
increased
urinary
volume.
Increased
levels
of
urinary
protein,
glucose,
and
potassium,
and
increased
activity
of
urinary
(­
glutamyltransferase
and
alkaline
phosphatase
were
also
observed
at
200
mg/
kg.
Histopathological
examination
of
the
kidneys
revealed
limited
focal
necrosis
at
100
mg/
kg
and
extensive
tubular
HCBD
 
February
2003
7­
3
necrosis
at
200
mg/
kg.
The
study
authors
identified
10
mg/
kg
and
100
mg/
kg
as
the
acute
"
No
Nephrotoxic­
Effect
Level"
and
"
Minimum
Nephrotoxic­
Effect
Level",
respectively.

Payan
et
al.
(
1993)
administered
single
oral
doses
of
0,
100,
or
200
mg/
kg
HCBD
in
polyethylene
glycol
to
male
Sprague­
Dawley
rats
(
4
to
5
animals
per
dose).
All
rats
were
sacrificed
24
hours
after
exposure,
and
the
right
kidneys
were
subjected
to
microscopic
examination.
Nephrotoxicity
was
also
evaluated
by
determination
of
the
following
biochemical
urinary
parameters
of
renal
impairment:
urinary
concentration
of
$
2­
microglobulin
($
2­
m),
urinary
(­
glutamyl
transpeptidase
(
 ­
GT)
activity,
urinary
aspartate
aminotransferase
(
AST)
activity,
and
urinary
Nacetyl
$­
glucosamine
(
NAG)
activity
At
100
mg/
kg,
4/
5
of
the
rat
kidneys
exhibited
mild
or
moderate
lesions,
compared
with
0/
4
in
the
controls.
Significant
increases
in
 
2­
m
(
2­
fold)
and
AST
(
19­
fold)
were
seen
at
100
mg/
kg,
and
in
all
biochemical
markers
of
renal
toxicity
at
200
mg/
kg.
A
Lowest­
Observed­
Adverse­
Effect
Level
(
LOAEL)
of
100
mg/
kg
was
identified
in
this
study
on
the
basis
of
kidney
lesions
and
a
increased
urinary
 
2­
m
and
AST
excretion.

Lock
et
al.
(
1996)
administered
a
single
oral
dose
of
50
mg
HCBD/
kg
to
a
calf
to
evaluate
toxic
effects
on
the
kidney
and
bone
marrow.
The
administered
dose
resulted
in
the
death
of
the
animal
5
days
after
treatment.
Prior
to
death,
blood
urea
nitrogen,
plasma
aspartate
aminotransferase
and
plasma
alkaline
phosphatase
were
elevated,
but
no
changes
were
observed
in
circulating
white
cells
or
platelets.
The
liver
and
kidneys
appeared
pale
and
swollen
at
necropsy.
Histopathological
examination
revealed
midzonal
necrosis
in
the
liver.
Extensive
areas
of
necrosis
were
evident
in
the
kidney,
and
were
accompanied
by
hyaline
and
granular
cast
formation.

Inhalation
Exposure
De
Ceaurriz
et
al.
(
1988)
evaluated
the
effects
of
HCBD
inhalation
exposure
on
male
Swiss
OF1
mice
(
6
mice/
dose).
The
mice
were
exposed
to
HCBD
vapor
at
concentrations
between
83
and
246
ppm
(
886
and
2,625
mg/
m3
)
for
15
minutes.
Decreased
respiratory
rates
(
reflex
bradypnea)
were
observed
at
concentrations
of
155
ppm
(
1,652
mg/
m3
)
or
greater.
An
EC
50
(
concentration
producing
an
effect
in
50%
of
the
population)
of
211
ppm
(
2,250
mg/
m3)
was
calculated
for
this
effect.

De
Ceaurriz
et
al.
(
1988)
investigated
the
effects
of
a
4­
hour
whole­
body
exposure
to
HCBD
at
measured
concentrations
of
2.75,
5,
10,
25
ppm
(
or
29.3,
53.4,
106.7,
266.8
mg/
m3)
on
male
Swiss
OF1
mice
(
10
animals/
dose).
An
HCBD­
related
increase
in
the
percentage
of
damaged
renal
tubules,
as
determined
by
alkaline
phosphatase
staining,
was
observed
at
all
exposure
levels.
The
EC
50
for
this
response
was
7.2
ppm
(
76.8
mg/
m3).

Gehring
and
MacDougall
(
1971)
exposed
rats
to
161
ppm
(
1,716
mg/
m3)
HCBD
for
0.88
hour
or
34
ppm
(
362
mg/
m3)
for
3.3
hours.
All
rats
survived
the
treatment.
Exposure
of
guinea
pigs
or
cats
under
the
same
conditions
resulted
in
the
death
of
most
animals.
Inhalation
exposure
of
rats
to
133
to
150
ppm
(
1,418
to
1,600
mg/
m3)
for
4
to
7
hours
was
lethal
for
some
or
all
animals
(
NTP,
1991).
HCBD
 
February
2003
7­
4
Dermal
Exposure
Gradiski
et
al.
(
1975)
evaluated
the
dermal
toxicity
and
sensitization
potential
of
HCBD
in
rabbits.
Dermal
application
of
a
10%
solution
of
HCBD
(
solvent
not
indicated)
to
rabbits
caused
slight
dermal
irritation.
Guinea
pigs
exhibited
delayed
allergic
reactions
to
dermal
HCBD
application
(
U.
S.
EPA,
1991a).

Duprat
and
Gradiski
(
1978)
evaluated
the
acute
toxicity
of
dermally
applied
HCBD
in
female
New
Zealand
rabbits
(
10
animals/
dose)
following
an
8­
hour
exposure
period.
Undiluted
HCBD
was
applied
at
doses
of
0.25,
0.5,
0.75
and
1.0
mL/
kg
under
occluded
conditions.
Four
hours
after
termination
of
exposure,
the
epidermis
and
subcutaneous
tissue
revealed
edema
and
polymorphonuclear
leukocyte
infiltration
at
the
two
highest
doses.
Three
to
five
days
after
treatment
at
the
three
highest
doses,
necrotic
changes
were
noted
at
the
site
of
application.
A
few
animals
from
the
two
highest
dose
groups
died
within
24
hours
from
respiratory
and
cardiac
failure.
Indications
of
systemic
toxicity
included
renal
epithelial
necrosis
and
fatty
liver
degeneration.
The
LD
50
for
the
eight
hour
exposure
was
0.72
mL/
kg
(
Duprat
and
Gradiski,
1978).
Based
on
a
specific
gravity
of
1.675
for
HCBD
(
U.
S.
EPA,
1991a),
a
dermal
LD
50
of
1,206
mg/
kg
was
calculated.

Acute
Ocular
Toxicity
Gradiski
et
al.
(
1975)
reported
that
instillation
of
a
10%
solution
of
HCBD
(
solvent
not
reported)
into
rabbit
eyes
resulted
in
slight
ocular
conjunctival
irritation
(
U.
S.
EPA,
1991a).

Intraperitoneal
Injection
Bai
et
al.
(
1992)
exposed
male
Sprague­
Dawley
rats
(
4
animals/
dose)
by
intraperitoneal
injection
0,
10.4,
52.2,
or
104
mg/
kg­
day
HCBD
for
three
days.
Serum
bilirubin
and
alkaline
phosphatase
activity
was
increased
(
p<
0.05)
at
the
two
higher
doses,
indicating
disturbance
in
liver
function.
The
concentration
of
total
serum
bile
acids
was
elevated
at
the
highest
dose.
No
histopathological
examination
of
the
livers
was
conducted.

Lock
and
Ishmael
(
1979)
administered
single
intraperitoneal
doses
of
HCBD
in
corn
oil
to
male
albino
rats
(
3
to
19
animals/
dose).
The
doses
administered
were
0,
20,
50,
100,
200,
300,
400,
500,
and
1,000
mg/
kg.
The
effects
of
treatment
were
assessed
24
hours
after
administration.
Three
of
four
animals
treated
with
500
mg/
kg
died.
All
rats
in
the
1,000
mg/
kg
dose
group
died.
Rats
in
the
highest
dose
group
exhibited
piloerection,
sedation,
hunching,
incoordination,
loss
of
muscle
tone
and
hypothermia
prior
to
death.

Wolf
et
al.
(
1983)
conducted
studies
to
examine
the
effects
of
HCBD
on
the
cytochrome
P450
content
and
related
monooxygenase
activities
in
the
kidneys
of
male
and
female
rats
and
male
mice.
Adult
male
and
female
Wistar
rats
were
administered
either
200
or
400
mg/
kg
HCBD
in
corn
oil
by
intraperitoneal
injection.
Male
Swiss­
derived
mice
were
administered
50
mg/
kg
HCBD
in
corn
oil
by
intraperitoneal
injection.
Control
animals
received
corn
oil
alone
(
5
ml/
kg).
A
subset
of
the
animals
were
treated
with
100
mg/
kg
 ­
naphthoflavone
(
BNF)
by
intraperitoneal
injection
daily
for
3
days
before
HCBD
administration.
Following
HCBD
administration,
animals
were
fasted
for
24
HCBD
 
February
2003
7­
5
hours,
and
renal
and
hepatic
microsomal
fractions
were
obtained.
Renal
cytochrome
P450
concentrations
dropped
significantly
(
p
<
0.01)
in
both
male
and
female
rats
at
both
doses
without
significant
losses
of
cytochrome
b5
or
NADPH­
cytochrome
c.
In
rat
renal
microsomes,
HCBD
increased
metabolism
of
aldrin
and
7­
ethoxycoumarin
significantly
(
p
<
0.05),
while
metabolism
of
p­
nitroanisole
was
only
slightly
reduced
and
metabolism
of
lauric
acid
was
unaffected.
No
effect
was
found
on
rat
hepatic
cytochrome
P450
levels,
and
no
reductions
in
rat
hepatic
monooxygenase
activity
were
found,
although
metabolism
of
7­
ethoxycoumarin
was
significantly
(
p
<
0.05)
increased.
In
BNF­
treated
rats,
HCBD
decreased
renal
cytochrome
P450
concentrations
33%,
decreased
metabolism
of
aldrin
significantly
(
p
<
0.05),
and
increased
metabolism
of
7­
ethoxycoumarin
and
7­
ethoxyresorufin.
In
naive
mice,
HCBD
treatment
led
to
an
almost
complete
loss
(
15%
of
control
remaining)
of
cytochrome
P450
and
associated
monooxygenase
activity
(
0­
28%
of
control
metabolism).
In
BNF­
treated
mice,
a
similar
loss
of
cytochrome
P450
(
17%
of
control
remaining)
resulted
in
similar
decreases
in
metabolism
(
0­
40%
of
control
metabolism),
except
for
metabolism
of
7­
ethoxycoumarin
(
67%
of
control
metabolism)
and
7­
ethoxyresorufin
(
117%
of
control
metabolism).
These
findings
suggest
that
the
cytochrome
P450
subtypes
which
are
decreased
are
localized
in
regions
where
HCBD
damage
occurs,
while
BNF­
induced
subtypes
are
not..

Hook
and
colleagues
(
Hook
et
al.,
1982,
1983;
Kuo
and
Hook,
1983)
conducted
a
series
of
experiments
to
characterize
HCBD
toxicity
in
four
different
strains
of
rats.
Single
doses
of
25
to
400
mg/
kg
HCBD
were
administered
by
intraperitoneal
injection.
Following
treatment,
relative
kidney
weight
increased
in
all
dose
groups.
Organic
ion
transport
was
evaluated
by
analysis
of
the
anion
paminohippurate
(
PAH)
and
the
cation
tetraethylammonium
(
TEA)
in
renal
cortical
slices.
Renal
efflux
rates
of
PAH
and
TEA
were
unaffected.
However,
accumulation
of
the
PAH
anion
was
decreased,
while
accumulation
of
the
TEA
cation
was
unaffected.
This
pattern
suggests
a
specific
impact
of
HCBD
on
the
renal
anion
uptake
system
(
Hook
et
al.,
1982;
Kuo
and
Hook,
1983).
Kidney­
to­
body
weight
ratios
increased
in
all
dosage
groups.
Blood
urea
nitrogen
levels
were
elevated
in
all
dose
groups,
with
a
more
pronounced
effect
noted
in
young
rats
(
Kuo
and
Hook,
1983).
Adult
male
rats
were
less
susceptible
to
HCBD­
induced
renal
effects
than
were
female
adult
rats
or
young
male
rats.
The
investigators
attributed
this
pattern
to
age­
and
sex­
related
differences
in
the
renal
and
hepatic
enzymes
responsible
for
activation
and
detoxification
of
HCBD.

Lock
et
al.
(
1984)
investigated
the
effect
of
age,
strain,
sex,
and
monooxygenase
inhibitors
on
the
acute
toxicity
of
HCBD
to
five
different
strains
of
mice
(
6
or
more
animals/
dose
group).
HCBD
in
corn
oil
was
administered
as
single
intraperitoneal
doses
ranging
from
6.3
to
50
mg/
kg.
Toxicity
was
evaluated
24
hours
after
treatment.
Histopathological
examination
of
the
kidneys
from
adult
Swiss­
derived
mice
of
both
sexes
revealed
extensive,
dose­
dependent
proximal
tubular
necrosis
at
doses
of
12.5
mg/
kg
and
above.
At
6.3
mg/
kg,
tubular
necrosis
was
only
observed
occasionally
in
a
small
number
of
animals.
A
significant
increase
in
plasma
urea
occurred
in
adult
Swiss­
derived
mice
at
doses
of
25
to
50
mg/
kg.
No
evidence
for
a
gender
difference
in
response
was
observed.
Young
male
mice
responded
to
lower
doses
of
HCBD
than
adults,
with
an
increase
in
plasma
urea
and
a
decrease
in
organic
ion
transport
in
renal
slices
evident
at
12.5
mg/
kg.
Prior
administration
of
the
monooxygenase
inhibitor
piperonyl
butoxide
or
the
monooxygenase
inducers
phenobarbitone
or
$­
naphthoflavone
did
not
modify
the
extent
of
HCBD­
induced
renal
damage.
Intraperitoneal
administration
of
the
glutathione
or
N­
acetylcysteine
conjugates
resulted
in
a
pattern
of
renal
necrosis
similar
to
that
observed
for
HCBD.
Evaluation
of
the
comparative
susceptibility
of
five
mouse
HCBD
 
February
2003
7­
6
strains
indicated
that
the
BALB/
c
strain
was
slightly
more
susceptible
to
HCBD
toxicity
than
the
C57BL/
10J,
C3H,
DBA/
2J,
and
Swiss­
derived
strains.

Ishmael
et
al.
(
1984)
investigated
the
time
course
of
histopathology
and
functional
impairment
in
adult
Swiss­
derived
mice
following
a
single
50
mg/
kg
intraperitoneal
dose
of
HCBD
in
corn
oil.
Histopathological
examination
of
the
kidneys
by
light
and
electron
microscopy
revealed
mitochondrial
swelling
within
1
hr
after
dosing,
followed
by
nuclear
pyknosis
and
increased
cytoplasmic
eosinophilia
at
4
hr,
extensive
tubular
necrosis
after
16
hr,
and
active
tubular
regeneration
within
5
days.
Treated
animals
exhibited
significantly
increased
plasma
urea,
decreased
renal
non­
protein
sulfhydryl
content,
and
increased
renal
water
content.

7.2.2
Short­
Term
Studies
Oral
exposure
Kociba
et
al.
(
1971)
conducted
a
study
of
HCBD
toxicity,
the
results
of
which
were
published
as
Kociba
et
al.
(
1977)
and
Schwetz
et
al.
(
1977).
Female
Sprague­
Dawley
rats
(
4
animals/
dose
group)
were
fed
diets
containing
HCBD
at
doses
of
0,
1,
3,
10,
30,
65,
or
100
mg/
kg­
day
for
30
days.
Renal
toxicity
in
the
form
of
increased
relative
kidney
weight
as
well
as
renal
tubular
degeneration,
necrosis
and
regeneration
was
observed
in
rats
receiving
doses
of
30,
65
or
100
mg/
kg­
day.
Minimal
hepatocellular
swelling
was
noted
at
a
dose
of
100
mg/
kg­
day.
Other
observed
effects
included
decreased
food
consumption,
reduced
body
weight
gain,
and
increased
hemoglobin
concentration
at
doses
of
10,
30,
65
or
100
mg/
kg­
day.
No
effects
were
observed
in
rats
receiving
3
mg/
kg­
day.
This
study
identified
a
No­
Observed­
Adverse­
Effect
Level
(
NOAEL)
of
3
mg/
kg­
day
and
a
LOAEL
of
10
mg/
kg­
day.

Harleman
and
Seinen
(
1979)
administered
diets
containing
nominal
concentrations
of
0,
50,
150,
or
450
ppm
HCBD
to
weanling
Wistar­
derived
rats
(
6
animals/
sex/
dose
group)
for
14
days.
Based
on
mean
body
weight
and
food
consumption
data
in
the
study,
the
mean
HCBD
doses
were
calculated
to
be
0,
4.6,
14.0
and
35.3
mg/
kg­
day.
Body
weight
and
food
conversion
efficiency
were
decreased
in
a
dose­
related
manner.
Food
consumption
was
decreased
at
35.3
mg/
kg­
day.
Relative
kidney
weights
were
significantly
increased
at
the
two
highest
dose
levels.
A
dose­
related
degeneration
of
renal
tubular
epithelial
cells
was
observed
in
all
treated
animals,
particularly
in
the
straight
limbs
(
pars
recta)
of
the
proximal
tubules
located
in
the
outer
medulla.
No
indications
of
liver
toxicity
were
observed.
The
low
dose
of
4.6
mg/
kg­
day
represented
the
LOAEL.

Stott
et
al.
(
1981)
conducted
an
oral
exposure
study
in
adult
male
Sprague­
Dawley
rats.
Five
animals
per
treatment
group
were
given
daily
doses
of
HCBD
(
0,
0.2
or
20
mg/
kg­
day)
in
corn
oil
for
three
weeks.
In
animals
exposed
to
20
mg/
kg­
day,
a
decrease
in
body
weight
gain
and
an
increase
in
relative
kidney
weight
were
observed.
Histopathological
examination
of
the
kidneys
revealed
damage
in
the
middle
and
inner
cortical
regions,
with
loss
of
cytoplasm,
nuclear
pyknosis,
increased
basophilia
and
mitotic
activity,
and
increased
cellular
debris.
No
indications
of
toxicity
were
observed
in
animals
exposed
to
0.2
mg/
kg­
day,
the
NOAEL
for
this
study.
HCBD
 
February
2003
7­
7
The
National
Toxicology
Program
(
NTP)
conducted
a
2­
week
oral
exposure
study
in
B6C3F
1
mice
(
NTP,
1991;
Yang
et
al.,
1989).
Groups
of
mice
(
5
animals/
sex/
dose
group)
received
diets
containing
nominal
concentrations
of
0,
30,
100,
300,
1,000
or
3,000
ppm
HCBD
for
15
days.
Target
concentrations
were
verified
under
experimental
conditions
by
gas
chromatography.
The
estimated
daily
intake
calculated
from
feed
consumption
and
mean
body
weights
were
0,
3,
12,
and
40
mg/
kgday
for
the
0,
30,
100,
and
300
ppm
dietary
concentrations,
respectively,
for
male
mice,
and
0,
5,
16,
and
49
mg/
kg­
day
for
female
mice.
All
mice
provided
with
the
1,000
and
3,000
ppm
diets
died
within
seven
days.
Mice
receiving
100
and
300
ppm
HCBD
lost
weight.
HCBD­
related
clinical
signs
observed
at
doses
of
300
ppm
or
greater
included
lethargy,
rough
hair
coats,
hunched
position,
and
incoordination.
Marked
reductions
in
thymus
and
heart
weights
were
noted
in
mice
consuming
the
300
ppm
diet.
Kidney
lesions
attributed
to
HCBD
exposure
were
observed
in
all
treated
mice
of
both
sexes
(
Yang
et
al.,
1989).
Severe
necrosis
of
the
cortex
and
outer
medulla
was
observed
in
the
two
doses
that
caused
deaths
among
the
experimental
animals.
Necrosis
was
less
severe
and
regeneration
was
prominent
in
the
pars
recta
of
mice
receiving
lower
doses
of
HCBD.
Histopathologic
changes
were
also
observed
in
liver,
lymphoid
tissues,
and
testis
of
mice
in
the
1,000
and
3,000
ppm
dose
groups,
but
were
not
clearly
related
to
HCBD
toxicity.
Minimal­
to­
mild
depletion
of
bone
marrow
in
the
femur
was
observed
in
2
to
5
mice
per
dose
group
in
animals
receiving
diets
containing
300
ppm
or
higher
levels
of
HCBD.
This
study
identified
a
LOAEL
of
3
 
5
mg/
kg­
day
in
male
and
female
mice,
respectively,
based
on
renal
tubular
necrosis
and
cellular
regeneration
in
animals
in
the
lowest
dose
groups
(
Yang
et
al.,
1989).

Jonker
et
al.
(
1993b)
investigated
the
toxicity
of
HCBD
in
10­
week­
old
male
and
female
Wistar
rats
(
5
rats/
sex/
dose).
HCBD
was
provided
in
the
diet
at
levels
of
0,
25,
100,
or
400
ppm
for
a
duration
of
four
weeks.
Based
on
mean
body
weight
and
food
intake
data
in
the
study,
these
concentrations
correspond
to
average
daily
doses
of
0,
2.25,
8,
and
28
mg/
kg­
day.
Treatment­
induced
signs
of
toxicity
were
observed
at
doses
of
8
and
28
mg/
kg­
day
in
both
sexes.
The
observed
signs
included
decreased
liver
weight,
tubular
cytomegaly,
decreased
plasma
creatinine,
decreased
body
weight
(
10%
in
males
and
15%
in
females),
and
decreased
adrenal
weight.
Increased
plasma
aspartate
aminotransferase
activity
and
bilirubin
were
observed
at
the
28
mg/
kg­
day
dose.
The
NOAEL
and
LOAEL
identified
from
this
study
were
2.25
and
8
mg/
kg­
day,
respectively.

Lock
et
al.
(
1996)
conducted
two
short­
term
experiments
on
the
effects
of
HCBD
in
calves.
Each
experiment
evaluated
toxicity
in
a
single
animal.
In
the
first
experiment,
a
dose
of
5
mg/
kg­
day
was
administered
orally
for
7
days.
An
increase
in
blood
urea
nitrogen
was
noted
after
the
fifth
dose,
and
levels
remained
high
until
the
animal
was
euthanized
nine
days
after
initiation
of
treatment.
Plasma
levels
of
aspartate
transaminase
and
alanine
aminotransferase
were
elevated,
but
no
changes
were
observed
in
hematological
parameters.
At
necropsy,
perirenal
edema
was
observed
in
the
kidneys
and
the
liver
was
pale
and
swollen.
Histopathological
examination
revealed
midzonal
necrosis
of
the
liver
and
extensive
swelling
of
the
tubular
epithelium
and
degenerative
changes
in
the
kidney.
Casts
were
evident
in
the
tubules
of
the
medulla.

In
the
second
short­
term
experiment
conducted
by
Lock
et
al.
(
1996),
a
calf
was
dosed
with
2.5
mg
HCBD/
kg­
day
for
10
days
and
the
blood
was
monitored
for
20
days
for
urea
and
platelet
count.
The
dose
was
subsequently
increased
to
5
mg/
kg­
day,
with
8
doses
administered
over
12
days.
A
marginal
increase
in
blood
urea
nitrogen
was
observed
on
day
14.
Aspartate
transaminase
and
HCBD
 
February
2003
7­
8
alanine
aminotransferase
were
increased
on
day
7,
and
gradually
decreased
to
normal
levels
on
day
15.
The
calf
was
euthanized
and
necropsied
18
days
after
the
start
of
the
dosing
regimen.
Histopathological
examination
revealed
slight
disruption
of
the
midzonal
architecture
of
the
liver,
while
mild
renal
tubular
degeneration
was
evident
in
the
kidney.
The
results
of
the
experiments
conducted
by
Lock
et
al.
(
1996)
indicate
that
HCBD
is
both
a
nephro­
and
hepato­
toxicant
in
calves.

Nakagawa
et
al.
(
1998)
exposed
male
Wistar
rats
(
3
animals/
dose
group)
to
0,
0.008,
0.04,
or
0.2%
HCBD
in
the
diet
for
3
weeks.
Assuming
a
food
consumption
factor
of
0.09
kg/
kg/
day
(
U.
S.
EPA,
1988),
these
dietary
levels
correspond
to
approximate
daily
doses
of
0,
7.2,
36,
and
180
mg/
kgday
Rats
ingesting
the
0.04%
and
0.2%
diets
had
lower
mean
body
weight
(
decreases
of
15%
and
46%,
respectively)
at
the
termination
of
the
experiment.
Kidney
weight
was
unaffected.
Histopathological
examination
of
rats
in
the
180
mg/
kg­
day
group
revealed
indications
of
extensive
regeneration
in
the
straight
portion
(
pars
recta)
of
the
proximal
tubule.
Similar
lesions
were
not
evident
at
lower
doses.
These
data
suggest
a
NOAEL
of
7.2
mg/
kg­
day
based
on
absence
of
effect
on
weight
gain
or
renal
histopathology.

Inhalation
exposure
NIOSH
(
1981)
reported
100%
mortality
in
mice
exposed
to
HCBD
vapors
for
5
days,
7
hours/
day,
at
a
concentration
of
50
ppm
(
533
mg/
m3),
but
no
deaths
in
animals
exposed
to
10
ppm
(
106.6
mg/
m3).

Gage
(
1970)
conducted
an
inhalation
study
in
Alderley
Park
SPF
rats.
Groups
of
adult
rats
(
4/
sex/
dose)
were
exposed
to
nominal
HCBD
concentrations
of
53,
107,
or
267
mg/
m3
for
15
days,
6
hours/
day
(
duration­
adjusted
concentrations
of
13,
27,
or
67
mg/
m3);
1,067
mg/
m3
for
12
days,
6
hours/
day
(
267
mg/
m3
duration­
adjusted);
or
2,668
mg/
m3
for
2
days,
4
hours/
day
(
445
mg/
m3
duration­
adjusted).
Petroleum
ether
was
used
as
a
solvent
for
concentrations
below
1,067
mg/
m3.
No
indications
of
toxicity
were
observed
at
the
lowest
level
of
exposure,
suggesting
a
NOAEL
of
53
mg/
m3
(
13
mg/
m3
duration­
adjusted).
Two
of
the
four
female
rats
exposed
to
1,067
mg/
m3
HCBD
died.
Pale
enlarged
kidneys,
adrenal
regeneration,
and
renal
cortical
tubular
degeneration
with
epithelial
regeneration
were
noted
at
autopsy.
Surviving
females
at
this
concentration
were
slightly
anemic.
The
weight
gain
of
female
rats
was
reduced
at
107
and
267
mg/
m3.
At
1,067
mg/
m3,
rats
of
both
sexes
lost
weight.
Irritation
of
the
eyes
and
nose
was
observed
at
the
two
highest
levels
of
exposure.
Respiratory
distress
occurred
at
concentrations
of
267
mg/
m3
or
greater.
At
the
termination
of
the
experiment,
enlarged
pale
kidneys
were
evident
in
the
267
and
1,067
mg/
m3
treatment
groups.
Enlarged
adrenals
were
observed
in
animals
exposed
to
1,067
mg/
m3.
Histopathological
analysis
revealed
degeneration
in
the
adrenal
cortex
and
proximal
tubules
of
the
kidneys
at
concentrations
of
267
mg/
m3
or
greater
(
WHO,
1994).

7.2.3
Subchronic
Studies
Schwetz
et
al.
(
1977)
fed
male
(
10
­
17
per
dose
group)
and
female
(
20
­
34
per
dose
group)
Sprague­
Dawley
rats
a
diet
containing
0.2,
2.0,
or
20
mg/
kg­
day
HCBD
for
evaluation
of
reproductive
effects.
HCBD
was
provided
in
the
diet
before
and
during
mating,
and
throughout
gestation
and
lactation,
for
a
total
study
duration
of
148
days.
Adult
rats
from
the
20
mg/
kg­
day
dose
HCBD
 
February
2003
7­
9
level
had
decreased
body
weight
gain
along
with
decreased
food
consumption.
At
necropsy,
relative
kidney
weights
were
increased
in
high­
dose
males
and
females.
Relative
liver
weight
was
increased
in
high­
dose
males,
and
relative
brain
weight
was
increased
in
high­
dose
females.
The
kidneys
of
males
exposed
to
2
(
1
of
10
examined)
or
20
(
3
of
15
examined)
mg/
kg­
day
were
roughened
and
had
a
mottled
cortex.
Histopathological
examination
revealed
dose­
related
increases
in
tubular
dilation
and
regeneration
in
animals
exposed
to
2
or
20
mg/
kg­
day
(
Table
7­
1).
These
results
indicate
a
NOAEL
of
0.2
mg/
kg­
day
for
male
and
female
rats,
based
on
the
absence
of
observed
renal
histopathology
or
other
toxic
effects
at
this
dose.

Table
7­
1.
Histopathological
Findings
in
Adult
Rats
Fed
Diets
containing
Hexachlorobutadiene.

Male
Rats
Female
Rats
Administered
Dose
(
mg/
kg­
day)
0
0.2
2
20
0
0.2
2
20
Moderate
Proteinaceous
Casts
in
Dilated
Renal
Tubules
1/
5
0/
5
3/
5
4/
5
0/
5
0/
5
1/
5
0/
5
Moderate
Focal
Renal
Tubular
Collapse
and
Atrophy
1/
5
0/
5
3/
5
4/
5
0/
5
0/
5
4/
5
0/
5
Dilation
and
Hypertrophy
of
the
Tubules
in
the
Kidney
Cortex
1/
5
0/
5
0/
5
4/
5
0/
5
0/
5
1/
5
4/
5
source:
Schwetz
et
al.
(
1977)

Harleman
and
Seinen
(
1979)
exposed
groups
of
weanling
Wistar­
derived
rats
(
10/
sex/
dose
group)
to
daily
oral
gavage
doses
of
0,
0.4,
1.0,
2.5,
6.3,
or
15.6
mg/
kg­
day
HCBD
in
a
peanut
oil
vehicle.
The
study
duration
was
13
weeks.
Reductions
in
body
weight
gain,
food
consumption,
and
food
utilization
efficiency
were
noted
in
the
two
highest
dose
groups.
Dose­
related
increases
in
relative
kidney
weights
were
noted
in
all
treatment
groups
of
male
mice,
and
in
females
administered
6.3
or
15.6
mg/
kg­
day.
Proximal
tubular
degeneration
was
noted
in
males
treated
with
doses
of
6.3
mg/
kg­
day
and
above,
and
in
females
treated
with
doses
of
2.5
mg/
kg­
day
and
above.
This
effect
was
characterized
by
hyperchromatic
nuclei,
hypercellularity,
vacuolation
and
focal
necrosis
of
renal
epithelial
cells,
and
a
diminished
brush
border.
Polyuria
and
decreased
urine
osmolarity
were
noted
in
females
receiving
doses
equal
to
or
greater
than
2.5
mg/
kg­
day,
and
in
males
receiving
15.6
mg/
kg­
day.
Relative
liver
weights
were
increased
in
females
at
15.6
mg/
kg­
day
and
in
males
at
6.3
and
15.6
mg/
kg­
day.
Histological
examination
of
the
liver
revealed
increased
cytoplasmic
basophilia
only
in
males
treated
with
6.3
mg/
kg­
day
and
above.
Relative
spleen
weights
were
increased
in
males
at
15.6
mg/
kg­
day,
and
in
females
at
the
two
highest
doses.
The
study
authors
identified
NOAEL
values
of
1.0
mg/
kg­
day
for
females
and
2.5
mg/
kg­
day
for
males.

NTP
(
1991)
conducted
a
13­
week
oral
exposure
study
in
B6C3F
1
mice.
Groups
of
mice
(
10
animals/
sex/
dose
group)
received
diets
containing
0,
1,
3,
10,
30,
or
100
ppm
HCBD.
Target
concentrations
were
verified
under
experimental
conditions
by
gas
chromatographic
analysis.
Based
HCBD
 
February
2003
7­
10
on
average
food
consumption
and
body
weight
data,
these
concentrations
corresponded
to
dose
levels
of
0,
0.1,
0.4,
1.5,
4.9
or
16.8
mg/
kg­
day
for
males
and
0,
0.2,
0.5,
1.8,
4.5
or
19.2
mg/
kg­
day
for
females.
No
HCBD­
related
clinical
signs
or
deaths
were
observed.
Food
consumption
of
treated
and
control
animals
was
similar.
Reduced
body
weight
gain
was
reported
in
males
exposed
to
diets
containing
30
or
100
ppm
HCBD
(
decreases
of
49%
and
56%,
respectively)
and
in
females
exposed
to
the
100
ppm
diet
(
47%).
Relative
kidney
weight
was
significantly
decreased
(
p<
0.01)
in
the
three
highest
dose
groups
of
males,
and
in
females
in
the
highest
dose
group.
High­
dose
males
also
exhibited
decreased
relative
heart
weight,
although
no
histologic
lesions
were
reported
in
this
organ.

Histopathological
changes
were
noted
in
the
kidneys
of
treated
animals.
Necropsy
revealed
treatment­
related
increases
in
renal
tubular
cell
regeneration.
This
lesion
was
characterized
as
a
diffuse
increase
in
epithelial
nuclei
and
increased
basophilic
staining.
Female
mice
appeared
to
be
more
susceptible
than
male
mice
to
the
formation
of
this
lesion
following
exposure
to
HCBD,
with
occurrence
noted
at
dose
levels
of
0.2
mg/
kg­
day
and
above.
Incidence
of
regeneration
in
female
mice
was
0%
at
control,
10%
at
0.2
mg/
kg­
day,
90%
at
0.6
mg/
kg­
day,
and
100%
at
higher
doses.
Lesions
were
observed
in
male
mice
at
dose
levels
of
4.9
mg/
kg­
day
and
above,
with
incidences
of
100%
in
these
groups
(
NTP,
1991;
Yang
et
al.,
1989).
In
contrast
to
results
from
the
2­
week
study
conducted
by
the
same
investigators,
no
evidence
of
necrosis
was
observed.
Based
on
the
histopathologic
evaluation
of
the
kidney,
the
authors
identified
a
NOAEL
of
1.5
mg/
kg­
day
for
male
mice.
Because
tubular
regeneration
occurred
in
1
of
10
females
in
the
lowest
dose
group
(
0.2
mg/
kgday
the
study
authors
concluded
that
a
NOAEL
for
female
mice
could
not
be
identified
from
these
data
(
NTP,
1991).
However,
others
have
concluded
that
the
effect
observed
at
0.2
mg/
kg­
day
is
not
statistically
significant,
and
therefore
considered
this
dose
to
be
the
NOAEL
for
female
mice
(
WHO,
1994;
U.
S.
EPA,
1998a).
Further
statistical
analysis
shows
that,
although
the
0.2
mg/
kg­
day
dose
fails
a
Fisher's
exact
test
against
control
(
p
=
0.5),
the
data
passes
a
Mantel­
Haenszel
trend
test
(
p
<
0.001),
suggesting
that
a
benchmark
dose
analysis
would
be
appropriate.

Nakagawa
et
al.
(
1998)
administered
0.1%
HCBD
in
the
diet
to
male
Wistar
rats
(
21
rats/
group)
for
30
weeks
in
conjunction
with
a
cancer
promotion
study
(
discussed
in
Section
7.2.7).
Assuming
a
food
consumption
factor
of
0.09
kg/
kg/
day
(
U.
S.
EPA,
1988),
this
dietary
level
corresponds
to
an
average
daily
dose
of
90
mg/
kg­
day.
HCBD
treatment
resulted
in
decreased
final
body
weight,
and
increased
relative
kidney
weight.
No
significant
differences
were
noted
in
serum
and
urine
biochemical
parameters.
Simple
hyperplasia
of
renal
tubular
structures
was
observed,
but
the
incidence
did
not
differ
significantly
from
the
control.
Histopathological
examination
did
not
reveal
adenomatous
hyperplastic
foci
or
renal
tumors.

7.2.4
Neurotoxicity
Data
from
distribution
studies
indicate
that
HCBD
accumulates
in
brain
tissue
(
Reichert
et
al.,
1985).
This
observation
raises
the
possibility
that
HCBD
exposure
may
affect
neurological
function.
Studies
designed
to
specifically
evaluate
neurotoxicological
endpoints
following
HCBD
exposure
were
not
identified
in
the
available
literature.
However,
neurological
effects
have
been
observed
in
a
number
of
oral
and
dermal
exposure
studies.
Kociba
et
al.
(
1977)
reported
increased
relative
brain
weights
in
female
rats
fed
20
mg/
kg­
day
HCBD
for
2
years.
This
increase
occurred
in
a
dose
group
with
depressed
body
weights
and
was
not
accompanied
by
histopathological
changes
HCBD
 
February
2003
7­
11
in
the
brain.
Similarly,
Schwetz
et
al.
(
1977)
noted
depressed
body
weight
and
increased
relative
brain
weight
in
female
rats
fed
20
mg/
kg­
day
for
148
days
in
a
reproductive
study.
Concurrent
changes
in
behavior
or
brain
histopathology
were
not
observed
in
the
affected
animals.
An
increase
in
relative
brain
weights
with
decreased
body
weights
was
also
observed
in
male
and
female
B6C3F
1
mice
fed
16.8
to
19.2
mg/
kg­
day
HCBD
in
their
diet
in
a
90­
day
subchronic
study
(
NTP,
1991).

Treatment­
associated
neurotoxic
effects
were
observed
by
Harleman
and
Seinen
(
1979),
who
provided
female
Wistar
rats
(
6
animals/
dose)
with
diets
containing
0,
150,
or
1,500
ppm
HCBD
(
corresponding
to
average
daily
doses
of
0,
15
or
150
mg/
kg­
day).
The
duration
of
exposure
ranged
from
10
to
18
weeks.
Indications
of
neurotoxicity
observed
at
the
150
mg/
kg­
day
dose
included
ataxia,
incoordination,
weakness
of
the
hind
legs,
and
unsteady
gait.
Histopathological
examination
revealed
demyelination
and
fragmentation
of
femoral
nerve
fibers
in
high­
dose
animals.
No
treatment­
related
histopathological
changes
were
observed
in
the
brain.

In
Russian
studies
cited
through
secondary
sources
(
WHO,
1994),
Badaeva
(
1983)
and
Badaeva
et
al.
(
1985)
observed
that
daily
oral
administration
of
8.1
mg/
kg­
day
HCBD
to
pregnant
rats
throughout
gestation
resulted
in
histopathological
changes
in
nerve
cells
and
myelin
fibers
of
the
spinal
cord
in
the
dams
and
their
offspring.
Increased
levels
of
free
radicals
were
detected
in
the
brain
and
spinal
cord
of
the
offspring
of
treated
dams
(
U.
S.
EPA,
1991a).

Neurotoxicity
has
also
been
observed
following
HCBD
exposure
by
the
dermal
route.
Duprat
and
Gradiski
(
1978)
observed
central
nervous
system
depression
manifested
as
stupor
in
rabbits
following
application
of
0.25
to
1.0
ml/
kg
(
418
to
1,675
mg/
kg)
in
an
acute
dermal
toxicity
test.
Stupor
was
observed
throughout
the
8­
hour
exposure
period,
and
during
a
2­
hour
period
immediately
following
exposure.

7.2.5
Developmental/
Reproductive
Toxicity
Oral
Exposure
Schwetz
et
al.
(
1977)
provided
male
and
female
Sprague­
Dawley
rats
(
30
to
51
animals/
dose
group)
with
a
diet
containing
HCBD
at
levels
corresponding
to
doses
of
0,
0.2,
2.0,
or
20
mg/
kg­
day
HCBD.
The
HCBD­
containing
diet
was
administered
for
90
days
prior
to
mating,
15
days
during
mating,
22
days
during
gestation,
and
21
days
during
lactation.
Adults
in
the
two
higher
dose
groups
exhibited
multiple
signs
of
toxicity,
including
decreased
food
consumption,
decreased
body
weight
gain,
and
renal
tubular
degeneration.
No
HCBD­
related
effects
on
pregnancy
rate,
time
to
delivery,
neonatal
survival,
neonate
sex
ratio,
weanling
histopathology,
or
incidence
of
neonatal
external,
visceral,
or
skeletal
anomalies
were
observed.
Slightly
decreased
pup
weight
(
p<
0.05)
was
observed
in
the
20
mg/
kg­
day
treatment
group
at
postnatal
day
21.
The
identified
NOAEL
and
LOAEL
for
developmental
effect
were
2
and
20
mg/
kg­
day,
respectively.

Harleman
and
Seinen
(
1979)
provided
female
rats
(
6
animals/
dose)
with
a
diet
containing
HCBD
at
levels
of
0,
150,
or
1,500
ppm
for
3
weeks
prior
to
mating,
3
weeks
during
mating,
and
throughout
gestation
and
lactation.
Assuming
a
food
consumption
factor
of
0.1
kg/
kg/
day
(
U.
S.
EPA,
1988),
these
concentrations
correspond
to
average
daily
doses
of
0,
15,
or
150
mg/
kg­
day.
1
Although
the
Guidelines
for
Developmental
Toxicity
Risk
Assessment
(
U.
S.
EPA,
1991b)
recommended
against
dosimetric
adjustment
of
developmental
toxicity
data,
the
draft
Review
of
the
Reference
Dose
and
Reference
Concentration
Processes
(
U.
S.
EPA,
2002)
recommends
that
duration
adjustment
procedures
to
continuous
exposures
be
used
for
inhalation
developmental
toxicity
studies
as
for
other
health
effects
from
inhalation
exposure.

HCBD
 
February
2003
7­
12
High­
dose
females
were
sacrificed
at
week
10
of
the
study,
and
low­
dose
females
were
sacrificed
at
week
18.
Maternal
toxicity
occurred
in
both
dose
groups,
and
observed
effects
included
reduced
body
weight
gain,
increased
relative
kidney
weight,
and
histopathological
changes
in
kidneys.
Neurological
effects,
including
ataxia,
incoordination,
weakness
of
the
hind
legs,
and
unsteady
gait,
were
observed
in
the
dams
at
the
150
mg/
kg­
day
dose.
Furthermore,
at
the
150
mg/
kg­
day
dose
no
conceptions
occurred,
the
ovaries
showed
little
follicular
activity,
and
no
uterine
implantation
sites
were
observed.
At
15
mg/
kg­
day,
fertility
and
litter
size
were
reduced,
but
the
effect
was
not
statistically
significant.
Pup
weights
in
this
treatment
group
were
significantly
reduced
on
postnatal
days
0,
10,
and
20.
No
gross
abnormalities
were
noted.
The
LOAEL
identified
was
15
mg/
kg­
day.

In
Russian
studies
cited
in
a
secondary
source
(
WHO,
1994),
Badaeva
and
colleagues
(
Badaeva,
1983;
Badaeva
et
al.,
1985)
conducted
two
studies
in
which
pregnant
female
rats
were
orally
administered
8.1
mg/
kg­
day
HCBD
throughout
gestation.
Offspring
of
HCBD­
treated
dams
had
lower
body
weight
and
shorter
crown­
rump
length
when
compared
to
controls
(
U.
S.
EPA,
1991a).
Histological
changes
in
the
nerve
cells
and
myelin
fibers
of
the
spinal
cord
were
noted
in
both
dams
and
offspring.
Neurological
changes
reported
in
the
offspring
included
ultrastructural
changes
in
neurocytes
and
increased
levels
of
free
radicals
in
the
brain
and
spinal
cord
(
U.
S.
EPA,
1991a).

In
addition
to
the
reproductive
and
developmental
toxicity
studies
discussed
above,
two
longer­
term
toxicity
studies
have
evaluated
reproductive
endpoints
following
oral
exposure
to
HCBD.
No
treatment­
related
lesions
in
reproductive
organs
were
observed
in
rats
that
received
lifetime
exposures
of
up
to
20
mg/
kg­
day
HCBD
(
Kociba
et
al.,
1977).
No
significant
changes
were
noted
in
sperm
count,
the
incidence
of
abnormal
sperm,
estrual
cyclicity,
or
the
average
estrous
cycle
length
in
mice
administered
100
ppm
HCBD
in
the
diet
for
13
weeks.
Sperm
motility
in
treated
mice
was
significantly
lower,
though
not
in
a
dose­
related
manner,
than
that
observed
for
controls
(
NTP,
1991).

Inhalation
Exposure
Saillenfait
et
al.
(
1989)
exposed
pregnant
Sprague­
Dawley
rats
(
24
animals/
dose)
to
HCBD
vapor
at
nominal
concentrations
of
0,
21,
53,
107,
or
160
mg/
m3
(
0,
2,
5,
10,
or
15
ppm)
for
6
hours/
day
from
gestation
days
(
GD)
6
 
20,
resulting
in
duration­
adjusted
concentrations1
of
0,
5,
13,
27,
or
40
mg/
m3.
Concentrations
were
monitored
by
gas
chromatography.
Animals
were
sacrificed
on
GD
21.
Decreased
body
weight
gain
was
occurred
in
dams
exposed
to
53
or
160
mg/
m3.
Body
weight
was
decreased
(
p<
0.01)
in
male
(
9.5%)
and
female
(
12.9%)
fetuses
in
the
160
mg/
m3
treatment
group.
Fetal
body
weight
was
unaffected
at
lower
doses.
The
mean
number
of
implantation
sites,
total
fetal
losses,
resorptions,
number
of
live
fetuses,
pregnancy
rate,
and
fetal
sex
ratio
were
HCBD
 
February
2003
7­
13
comparable
in
the
treated
and
control
groups.
No
exposure­
related
external,
visceral,
or
skeletal
anomalies
were
noted
in
any
dose
group.
The
NOAEL
for
maternal
toxicity
is
at
5
mg/
m3.

In
dominant
lethal
tests
in
CD
(
Sprague­
Dawley­
derived)
rats,
exposure
to
HCBD
vapors
at
10
or
50
ppm
(
107
or
533
mg/
m3)
for
5
consecutive
days,
7
hours/
day,
did
not
affect
fertility
indices,
number
of
corpora
lutea
or
implantations,
or
the
frequency
of
early
death
(
NIOSH,
1981).

For
B6C3F
1
mice
that
were
exposed
to
HCBD
vapors
at
107
or
533
mg/
m3,
all
animals
in
the
high­
dose
group
died
during
the
5­
week
post­
treatment
period
(
NIOSH,
1981).
The
frequency
of
abnormal
sperm
morphology
in
the
low­
dose
group
did
not
differ
significantly
from
controls.

Intraperitoneal
Injection
Mated
female
Sprague­
Dawley
rats
(
10
 
15
animals/
group)
received
10
mg/
kg­
day
HCBD
in
corn
oil
via
intraperitoneal
injection,
during
gestation
days
1
to
15
(
Hardin
et
al.,
1981).
Maternal
toxicity
consisted
of
changes
in
two
organ
weights
(
no
further
details
provided).
Decreased
pre­
and
post­
implantation
survival
was
also
noted.
Developmental
effects
included
decreased
fetal
weight
or
length,
a
1­
to­
2­
day
delay
in
heart
development,
and
dilated
ureters.
Gross
external
and
internal
examinations
revealed
no
malformations
(
WHO,
1994).

Harris
et
al.
(
1979)
exposed
pregnant
female
rats
to
10
mg/
kg­
day
HCBD
from
gestation
days
1
to
15
via
intraperitoneal
injection.
A
3­
fold
increase
in
soft
tissue
anomalies
was
reported
in
offspring
of
treated
dams.
No
particular
type
of
anomaly
predominated
(
U.
S.
EPA,
1991a).

7.2.6
Chronic
Toxicity
Data
from
a
single
chronic
oral
exposure
study
are
available
in
the
published
literature.
Kociba
et
al.
(
1977)
provided
male
and
female
Sprague­
Dawley
rats
(
39
to
40/
sex/
dose
level;
90/
sex
for
controls)
with
diets
that
contained
0,
0.2,
2,
or
20
mg/
kg­
day
HCBD
(
99%
pure)
for
22
months
(
males)
or
24
months
(
females).
Parameters
monitored
included
appearance
and
demeanor,
body
and
organ
weights,
food
consumption,
hematologic
and
urine
analysis,
urinary
porphyrins,
serum
clinical
chemistry,
and
histopathology
of
major
organs.
The
investigators
reported
significantly
increased
mortality
in
high­
dose
males
(
p<
0.05),
but
incidences
were
not
given.
Decreased
body
weight
gain
was
noted
in
high­
dose
males
and
females,
with
significant
differences
(
p<
0.05)
from
controls
evident
after
day
27
(
females)
or
day
69
(
males).
There
were
no
apparent
treatment­
related
effects
on
food
consumption.
High­
dose
animals
had
increased
relative
brain
weights
(
females)
and
relative
testes
weights
(
males).

An
important
observation
in
the
Kociba
et
al.
(
1977)
study
is
the
clear
dose­
response
relationship
for
HCBD­
induced
renal
toxicity.
No
discernible
effects
were
noted
at
the
0.2
mg/
kgday
dose.
Effects
noted
at
the
intermediate
dose
of
2
mg/
kg­
day
included
increased
urinary
coproporphyrin
excretion
(
females
only,
days
427­
428),
and
increased
renal
tubular
epithelial
hyperplasia.
Lifetime
ingestion
of
the
20
mg/
kg­
day
dose
resulted
in
increased
urinary
excretion
of
coproporphyrin
and
increased
terminal
weight
of
the
kidneys
in
rats
of
both
sexes
(
females
­
days
728­
729,
males
­
days
377­
378).
The
physiological
relationship
of
a
change
in
coproporphyrin
HCBD
 
February
2003
7­
14
excretion
to
renal
toxicity
is
not
currently
clear.
Microscopic
examination
revealed
histopathological
changes
in
the
kidney,
including
hyperplasia
and
neoplasia
of
the
renal
epithelium.
No
incidence
data
was
provided
for
the
non­
neoplastic
effects.
HCBD­
related
neoplastic
changes
are
further
discussed
in
Section
7.2.7.
The
lowest
dose
of
0.2
mg/
kg­
day
was
identified
as
the
NOAEL
in
this
study.
The
LOAEL
was
2
mg/
kg­
day.

7.2.7
Carcinogenicity
Oral
Exposure
Kociba
et
al.
(
1977)
observed
the
tumorigenic
potential
of
HCBD
in
male
and
female
rats
fed
0,
0.2,
2,
or
20
mg/
kg­
day
in
a
2­
year
oral
exposure
study.
No
adverse
effects
attributable
to
HCBD
were
observed
in
the
low­
dose
group.
Ingestion
of
the
intermediate
2
mg/
kg­
day
dose
resulted
in
signs
of
renal
tubular
epithelial
hyperplasia,
but
no
evidence
of
neoplasia
was
observed.
Ingestion
of
20
mg/
kg­
day
for
2
years
resulted
in
development
of
renal
tubular
adenomas
and
adenocarcinomas.
Neoplasms
were
observed
in
approximately
23%
(
9/
39)
of
the
males
and
15%
(
6/
40)
of
the
females
at
20
mg/
kg­
day,
compared
with
1.1%
(
1/
90)
and
0%
(
0/
90)
in
control
male
and
female
rats
and
0%
at
0.2
or
2
mg/
kg­
day.
Of
the
observed
neoplasms,
7
in
the
high
dose
males
and
3
in
the
high
dose
females
were
malignant,
the
rest
being
benign.
Combined
incidence
of
renal
tubular
benign
and
malignant
tumors
was
significantly
increased
when
compared
to
controls
(
p<
0.05)
for
both
males
and
females.
Metastasis
to
the
lungs
was
observed
in
two
of
the
treated
animals.
An
important
observation
in
this
study
was
that
HCBD­
induced
neoplasms
occurred
only
at
a
dosage
level
that
caused
substantial
renal
tissue
injury.
Additional
details
of
this
study
are
provided
in
Section
7.2.6.

In
a
Russian
study
cited
through
secondary
sources
(
USEPA,
1991a),
Chudin
et
al.
(
1985)
conducted
an
HCBD
oral
exposure
study
in
male
Wistar
rats
for
approximately
1
year.
The
doses
of
HCBD,
administered
by
gavage
in
sunflower
oil,
were
0.6,
5.8
or
37
mg/
kg­
day
(
n=
43,
43,
and
41,
respectively).
Control
rats
were
either
untreated
(
n=
90)
or
received
only
the
sunflower
oil
vehicle
(
n=
46).
Some
benign
tumors
of
the
kidney
and
liver
were
reported.

Nakagawa
et
al.
(
1998)
investigated
the
effect
of
HCBD
on
renal
carcinogenesis
in
male
Wistar
rats
pre­
treated
with
N­
ethyl­
N­
hydroxyethylnitrosamine
(
EHEN).
EHEN
is
a
known
nephrocarcinogen
in
rats,
where
it
selectively
induces
renal
tubular
cell
tumors.
The
purpose
of
this
study
was
to
evaluate
the
ability
of
HCBD
to
act
as
a
promoting
stimulus
following
subthreshold
exposure
to
EHEN.
HCBD
was
administered
for
30
weeks
at
a
concentration
of
0.1%
by
weight
(
dose
calculated
to
be1,000
mg/
kg)
in
the
diet
of
rats
(
12/
treatment
group)
that
had
previously
received
0.1%
EHEN
in
the
drinking
water.
The
combined
treatment
with
HCBD
and
EHEN
resulted
in
a
significantly
higher
renal
tumor
incidence
than
did
administration
of
EHEN
alone.
Rats
treated
with
HCBD
alone
did
not
develop
renal
tumors
under
the
conditions
used
in
this
investigation.
Significantly
increased
levels
of
bromodeoxyuridine
(
BrdU)
labeling
indicated
increased
cell
proliferation
in
the
outer
stripe
and
cortex
of
kidneys
from
HCBD­
treated
rats.
In
a
parallel
experiment,
immunostaining
for
proliferating
cell
nuclear
antigen
(
PCNA)
was
used
to
estimate
nuclear
DNA
synthesis
in
defined
renal
tubular
segments
of
HCBD­
treated
rats.
A
significant
increase
in
the
number
of
PCNA­
positive
cells
was
noted
only
in
the
outer
stripe.
These
results
are
HCBD
 
February
2003
7­
15
consistent
with
the
outer
stripe
as
a
site
for
renal
lesions
induced
by
HCBD.
Nakagawa
et
al.
(
1998)
concluded
that
the
ability
of
HCBD
to
induce
EHEN­
initiated
carcinogenesis
appears
to
be
associated
with
nephropathy
and
subsequent
cell
proliferation.

Dermal
Exposure
Van
Duuren
et
al.
(
1979)
evaluated
the
carcinogenicity
of
dermally
applied
HCBD
in
female
Swiss
mice
(
30
animals/
group).
The
investigators
applied
6.0
mg
of
HCBD
in
acetone
to
shaved
dorsal
skin
three
times
per
week
for
a
duration
of
440
to
594
days.
The
treatment
did
not
increase
the
incidence
of
papillomas
or
carcinoma
at
the
site
of
application,
or
the
incidence
of
tumors
at
distant
sites
such
as
the
lung,
stomach
or
kidney.

Van
Duuren
et
al.
(
1979)
also
evaluated
HCBD
in
an
initiation­
promotion
experiment.
Female
Swiss
mice
(
30
animals/
group)
received
a
single
application
of
15.0
mg
HCBD
in
acetone
on
shaved
dorsal
skin.
Fourteen
days
after
HCBD
application,
dermal
applications
of
5
µ
g
of
the
tumor
promoter
12­
o­
tetradecanoylphorbol­
13­
acetate
(
TPA)
were
administered
to
the
test
site
three
times
per
week
for
a
total
duration
of
428
to
576
days.
The
incidence
of
skin
papillomas
in
HCBDtreated
animals
was
comparable
to
that
in
controls.

Intraperitoneal
Injection
Theiss
et
al.
(
1977)
investigated
the
carcinogenic
potential
of
HCBD
by
assessment
of
the
pulmonary
tumor
response
in
male
strain
A/
St
mice.
Twenty
animals
per
dose
group
were
given
intraperitoneal
injections
of
4
or
8
mg/
kg
HCBD
in
tricaprylin,
three
times
per
week
for
a
total
of
13
and
12
injections,
respectively.
The
total
injected
dose
was
52
or
96
mg
HCBD
per
animal.
All
surviving
animals
were
killed
24
weeks
after
the
first
injection,
and
were
examined
for
pulmonary
surface
adenomas.
The
tumor
incidences
were
similar
in
treated
and
control
groups.
However,
the
use
of
this
study
for
the
evaluation
of
the
carcinogenicity
of
HCBD
is
limited
by
the
use
of
a
mouse
strain
that
is
highly
predisposed
to
spontaneous
lung
cancer,
the
small
number
of
animals
per
dose
group,
the
parenteral
route
of
administration,
the
limited
scope
of
histopathological
evaluation
(
WHO,
1994),
and
short
study
duration
(
6
months).

7.3
Other
Key
Data
7.3.1
Mutagenicity/
Genotoxicity
The
mutagenicity
of
HCBD
has
been
evaluated
in
an
array
of
in
vivo
and
in
vitro
assays.
The
results
of
these
tests
are
summarized
below
by
category.
No
information
was
located
regarding
the
genotoxic
effects
of
HCBD
in
humans.

Bacterial
Test
Systems
Test
results
from
bacterial
assays
of
mutagenicity
are
summarized
in
Table
7­
2.
Most
tests
in
standard
S.
typhimurium
reverse
mutation
assays
have
been
negative,
with
or
without
S9
activation
(
Rapson
et
al.,
1980;
Reichert
et
al.,
1983;
Stott
et
al.,
1981;
DeMeester
et
al.,
1981;
Haworth
et
al.,
HCBD
 
February
2003
7­
16
1983;
Chudin
et
al.,
1985)
except
for
a
study
by
Simmon
(
1977)
in
which
a
positive
response
in
Salmonella
typhimurium
was
reported
in
the
presence
of
metabolic
activation
induced
by
rat
liver
S9
fraction
(
U.
S.
EPA,
1991a).
Conflicting
results
in
standard
assays
may
be
due
to
contaminants
in
technical
and
even
analytical
grade
HCBD
(
Reichert
et
al.,
1984;
Vamvakas
et
al.,
1988).
Vamvakas
et
al.
(
1988)
observed
98%
pure
HCBD
was
a
direct­
acting
mutagen
in
S.
typhimurium
TA
100;
but
after
HCBD
was
purified
to
99.5%,
a
negative
mutagenic
response
was
obtained.

Positive
results
have
been
reported
for
HCBD
when
pre­
incubated
with
both
rat
liver
microsomes
and
glutathione,
but
not
when
either
was
omitted
(
Vamvakas
et
al.,
1988;
Roldan­
Arjona
et
al.,
1991).
The
mutagenic
response
was
increased
by
additional
inclusion
of
rat
kidney
microsomes,
mitochondria,
or
cytosol
as
(­
glutamyl
transpeptidase
and
dipeptidase
sources,
and
reduced
by
addition
of
a
 ­
lyase
inhibitor
(
Vamvakas
et
al.,
1988).
Reichert
et
al.
(
1984)
also
reported
a
positive
response
in
S.
typhimurium
when
a
"
fortified"
S9
mix,
containing
3­
fold
more
S9
protein
than
standard
Ames
test
protocols,
was
utilized.

Test
results
for
mutagenicity
assays
of
HCBD
metabolites
are
summarized
in
Table
7­
3.
Positive
results
in
bacterial
reverse
mutation
assays
have
been
obtained
for
the
mono­
glutathione
and
mono­
cysteine
conjugates
of
HCBD
(
Green
and
Odum,
1985;
Dekant
et
al.,
1986;
Vamvakas
et
al.,
1988)
and
the
mercapturic
acid,
N­
acetyl­
S­
pentachlorobutadienyl­
L­
cysteine
(
Reichert
and
Schutz,
1986;
Wild
et
al.,
1986).
Other
potential
HCBD
metabolites
that
gave
positive
results
in
reverse
mutation
assays
were
pentachloro­
3­
butenoic
acid
and
pentachloro­
3­
butenoic
acid
chloride
(
Reichert
et
al.,
1984),
in
the
absence
of
S9
activation.
When
rat
kidney
fractions
were
used
for
metabolic
activation,
the
addition
of
a
specific
inhibitor
of
$­
lyase
(
aminooxyacetic
acid)
to
the
system
reduced
the
mutagenic
response
(
Vamvakas
et
al.,
1988),
indicating
that
HCBD
metabolites
mediated
mutagenesis
in
these
assays.
These
results
are
consistent
with
the
proposed
mechanism
for
bioactivation
of
HCBD
in
animals
(
Figure
6­
1,
Section
6.3).

In
Vitro
Mammalian
Cell
Test
Systems
Data
for
mutagenicity
assays
in
mammalian
test
systems
are
summarized
in
the
upper
portion
of
Table
7­
4.
Treatment
with
HCBD
did
not
increase
the
frequency
of
chromosome
aberrations
in
Chinese
hamster
ovary
(
CHO)
cells
(
Galloway
et
al.,
1987)
or
cultured
human
lymphocytes
(
German,
1988).
However,
Galloway
et
al.
(
1987)
observed
a
significant
increase
in
sister
chromatid
exchange
in
CHO
cells
treated
with
HCBD.
Schiffman
et
al.
(
1984)
reported
unscheduled
DNA
synthesis
(
UDS)
activity
and
morphological
transformation
in
Syrian
hamster
embryo
fibroblasts
with
and
without
metabolic
activation.
Stott
et
al.
(
1981)
reported
negative
results
in
a
rat
primary
hepatocyte
UDS
assay.

Vamvakas
et
al.
(
1989)
evaluated
the
genotoxicity
of
the
HCBD
metabolite
S­(
1,2,3,4,4­
pentachlorobutadienyl)
glutathione
(
PCBG)
in
cultured
porcine
kidney
LLC­
PK
cells.
Incubation
of
confluent
monolayers
with
PCBG
resulted
in
a
dose­
dependent
induction
of
DNA
repair.
Addition
of
either
acivicin,
an
inhibitor
of
(­
glutamyl
transpeptidase,
or
aminooxyacetic
acid,
an
inhibitor
of
cysteine
conjugate
$­
lyase,
prevented
PCBG­
induced
genotoxicity.
These
results
are
consistent
with
the
hypothesis
that
renal
metabolism
plays
a
key
role
in
PCBG­
induced
genotoxicity.
HCBD
 
February
2003
7­
17
Table
7­
2.
Mutagenicity
of
HCBD
in
Salmonella
typhimurium
Test
Systems.

Strain
Conditions
Resultsb
Reference
With
S9
Activation
Without
S9
Activation
TA100
TA
1535
­
+
+
Simmon
(
1977)

TA100
purity
not
reported
nd
­
Rapson
et
al.
(
1980)

TA100
purity
>
99%
­
­
Stott
et
al.
(
1981)

TA98
TA100
TA1530
TA1535
TA1538
purity
98%,
plate
incorporation
­
­
De
Meester
et
al.
(
1981)

TA100
purity
>
99%,
plate
incorporation
­
­
Stott
et
al.
(
1981)

TA98
TA100
purity
not
reported,
suspension
test
­
­
Reichert
et
al.
(
1983)

TA98
TA100
TA1535
TA1537
purity
not
reported,
preincubation
test
­
­
Haworth
et
al.
(
1983)

+
fortified
S9*
a,

purity
>
99.5%,
preincubation
test
+
nd
TA100
TA1535
TA1538
+/­
rat
liver
S9,
purity
not
reported,
plate
incorporation
­
­
Chudin
et
al.
(
1985)

TA100
+/­
rat
liver
S9,
preincubation
test
­
­
Reichert
et
al.
(
1984)

+
rat
liver
S9*,
preincubation
test
+
­

TA100
+
rat
liver
S9*,
purity
99.5%,
preincubation
test
+
c
nd
Wild
et
al.
(
1986)
Table
7­
2
(
continued)

Strain
Conditions
Resultsb
Reference
With
S9
Activation
Without
S9
Activation
HCBD
 
February
2003
7­
18
TA100
purity
99.5%
nd
+
Vamvakas
et
al.
(
1988)

purity
99.5%,
preincubation
test
nd
­

TA100
+
rat
liver
microsomes
without
additional
GSH
­
nd
Vamvakas
et
al.
(
1988)

+
rat
liver
microsomes
and
additional
GSH,

purity
99.5%,

plate
incorporation
+
d
nd
TA100
purity
98%,
preincubation
test
­
+
Roldan­
Arjona
et
al.
(
1991)

Source:
adapted
from
WHO
(
1994)

a
S9*
=
a
fortified
S9
mix
containing
3
times
the
normal
protein
concentration;
GSH
=
reduced
glutathione
b
+
=
$
twice
the
background
rate
or,
in
the
case
of
bacterial
studies,
a
reproducible
dose­
related
increase
in
the
number
of
revertants
per
plate;
­
=
negative;
nd
=
not
determined
c
0.23
revertants
per
nmol
d
Addition
of
rat
kidney
microsomes
further
increased
the
number
of
revertants;
positive
results
were
inhibited
by
the
$­
lyase
inhibitor
aminooxyacetic
acid
HCBD
 
February
2003
7­
19
Table
7­
3.
Mutagenicity
of
HCBD
Metabolites.

Metabolitea
Conditions
Resultc
Reference
With
S9
Activation
Without
S9
Activation
PCBG
no
activation
nd
­
Green
&
Odum
(
1985)

+
rat
kidney
S9
+
nd
Green
&
Odum
(
1985)

+/­
rat
kidney
fractions
+
d
+
d
Vamvakas
et
al.
(
1988)

TCBG
+/­
rat
kidney
fractions
­
­
Vamvakas
et
al.
(
1988)

TCBC
+/­
rat
kidney
fractions
­
­
Vamvakas
et
al.
(
1988)

PCBC
+/­
rat
kidney
S9
+
e
+
e
Green
&
Odum
(
1985)

no
activation
nd
+
f
Dekant
et
al.
(
1986)

N­
AcPCBC
­
rat
liver
S9
nd
­
Wild
et
al.
(
1986)

+
rat
liver
S9
+
nd
+
rat
liver
S9
+
nd
Reichert
and
Schutz
(
1986)

PCCMT
+
rat
liver
S9*
b
­
nd
Wild
et
al.
(
1986)

PCMT
+
rat
liver
S9
­
nd
Wild
et
al.
(
1986)

PCBA
+
rat
liver
S9
+
+
Reichert
et
al.
(
1984)

PCBAC
+
rat
liver
S9
+
+
Reichert
et
al.
(
1984)

Source:
modified
from
WHO
(
1994)

aAbbreviations:
PCBG
S­(
1,1,2,3,4­
pentachlorobutadienyl)
glutathione
TCBG
1,4
bis(
1,2,3,4­
tetrachlorobutadienyl)
glutathione
TCBC
1,4
bis
(
1,2,3,4­
tetrachlorobutadienyl)­
L­
cysteine
PCBC
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine
N­
AcPCBC
N­
acetyl­
S­(
1,1,2,3,4­
pentachlorobutadienyl)­
L­
cysteine
PCCMT
1,1,2,3,4­
pentachlorobutadiene
carboxymethylthioether
PCMT
1,1,2,3,4­
pentachlorobutadiene
methylthioether
PCBA
2,2,3,4,4­
pentachloro­
3­
butenoic
acid
PCBAC
2,2,3,4,4­
pentachloro­
3­
butenoic
acid
chloride
b
S9*
=
a
fortified
S9
mix
containing
3
times
the
normal
protein
concentration;
GSH
=
reduced
glutathione
c
+
=
twice
background
rate;
­
=
negative;
nd
=
not
detectable
d
Mutagenic
potency
enhanced
by
rat
kidney
microsomes
or
mitochondria
and
less
so
by
cytosol;
positive
results
e
Mutagenic
potency
enhanced
by
rat
kidney
microsomes
or
mitochondria
and
less
so
by
cytosol;
positive
results
f
Mutagenic
potency
enhanced
by
the
 ­
lyase
inhibitor
aminooxyacetic
acid
HCBD
 
February
2003
7­
20
Table
7­
4.
Genotoxicity
of
HCBD
in
Eukaryotic
Assay
Systems.

Species/
Strain/

Cell
Type
Compound
End
Point
Comments
Resultsa
Reference
In
Vitro
Assays
Chinese
hamster
ovary
cells
HCBD
Chromosome
aberrations
+/­
rat
liver
S9
­
Galloway
et
al.
(
1987)

Human
lymphocytes
HCBD
Chromosome
aberrations
+/­
rat
liver
S9
­
German
(
1988)

Chinese
hamster
ovary
cells
HCBD
Sister
chromatid
exchange
+/­
rat
liver
S9
+
Galloway
et
al.
(
1987)

Syrian
hamster
embryo
fibroblast
HCBD
Unscheduled
DNA
synthesis
+/­
activation
+
Schiffmann
et
al.
(
1984)

Porcine
kidney
cells
PCBG
Unscheduled
DNA
synthesis
Addition
of
acivicin
or
aminooxyacetic
acid
gave
negative
results
+
Vamvakas
et
al.
(
1989)

Syrian
hamster
embryo
fibroblasts
HCBD
Morphological
transformation
+/­
activation
+
Schiffmann
et
al.
(
1984)

Rat
primary
hepatocyte
HCBD
Unscheduled
DNA
synthesis
­
Stott
et
al.
(
1981)

In
Vivo
Assays
Mouse;
bone
marrow
cells
HCBD
Chromosome
aberrations
Inhalation,
4h
+
German
(
1988)

Mouse;
bone
marrow
cells
HCBD
Chromosome
aberrations
Oral
gavage
+
German
(
1988)

Rat;
bone
marrow
cells
HCBD
Chromosomal
aberration
Dietary
doses
of
up
to
20
mg/
kg­
day
for
148
days
­
Schwetz
et
al.
(
1977)
Table
7­
4
(
continued)

Species/
Strain/

Cell
Type
Compound
End
Point
Comments
Resultsa
Reference
HCBD
 
February
2003
7­
21
Rat;
bone
marrow
cells
HCBD
Chromosomal
aberrations
10
or
50
ppm,
7
hrs/
day,

1
or
5
days
­
NIOSH
(
1981)

Rat
HCBD
Dominant
lethality
10
or
50
ppm,
7
hrs/
day,

5
days
­
NIOSH
(
1981)

Rat
kidney
cells
HCBD
DNA
alkylation
20
mg/
kg
by
gavage
+
Stott
et
al.
(
1981)

Rat
kidney
cells
HCBD
DNA
repair
20
mg/
kg
by
gavage
+
Stott
et
al.
(
1981)

Mouse
HCBD
DNA
binding
Single
dose,
30
mg/
kg
+
b
Schrenk
and
Dekant
(
1989)

Drosophila
melanogaster
HCBD
Gene
mutation
(

sexlinked
recessive
lethal)
­
NIOSH
(
1981)

Drosophila
melanogaster
HCBD
Sex­
linked
lethals
Feeding
or
injection
­
Woodruff
et
al.
(
1985)

Source:
modified
from
ATSDR
(
1994)
and
WHO
(
1994).

a
+
=
$
twice
the
background
rate
or,
in
the
case
of
bacterial
studies,
a
reproducible
dose­
related
increase
in
the
number
of
revertants
per
plate;
­
=
negative;

nd
=
not
determined;

b
Binding
was
predominately
to
mitochondrial
DNA
HCBD
 
February
2003
7­
22
In
Vivo
Test
Systems
Results
of
in
vivo
HCBD
genotoxicity
tests
are
summarized
in
the
lower
portion
of
Table
7­
4.
Both
negative
(
NIOSH,
1981;
Schwetz
et
al.,
1977)
and
positive
(
German,
1988)
results
have
been
reported
for
chromosome
aberration
assays
conducted
in
HCBD­
treated
mice
or
rats.
Negative
findings
have
been
reported
in
a
dominant
lethal
assay
in
rats
(
NIOSH,
1981),
and
in
the
Drosophila
melanogaster
sex­
linked
recessive
lethal
mutation
assay
with
exposure
via
either
injection
or
feeding
(
NIOSH,
1981;
Woodruff
et
al.,
1985).
However,
Stott
et
al.
(
1981)
reported
a
small
(
1.25
to
1.54­
fold)
increase
in
UDS
activity
and
DNA
alkylation
(
0.78
alkylation
per
106
nucleotides)
in
kidney
cells
from
rats
fed
20
mg/
kg­
day
HCBD
in
the
diet
for
3
weeks,
suggesting
that
HCBD
exhibited
a
minor
degree
of
renal
genotoxicity.

Schrenk
and
Dekant
(
1989)
evaluated
the
covalent
binding
of
HCBD
metabolites
to
renal
and
hepatic
DNA
in
NMRI
mice.
A
low
level
of
covalent
binding
(
covalent
binding
index
(
CBI)
=
27)
was
observed
with
nuclear
DNA
(
nDNA)
isolated
from
the
kidney,
while
covalent
binding
was
undetectable
in
nDNA
isolated
from
liver.
Significantly
higher
levels
of
covalent
binding
were
observed
with
mitochondrial
DNA
(
mtDNA),
with
CBIs
of
513
and
7,506
determined
for
liver
and
kidney,
respectively.
Analysis
of
covalent
binding
to
renal
mtDNA
identified
three
14C­
labeled
compounds
that
appeared
to
be
DNA
bases
altered
by
HCBD
metabolites.

7.3.2
Immunotoxicity
The
immunological
effects
of
HCBD
have
not
been
systematically
evaluated
in
humans,
and
there
are
currently
no
case
reports
that
describe
immunological
abnormalities
occurring
in
humans
exposed
to
HCBD.

Animal
data
on
the
immunological
effects
of
HCBD
are
limited.
In
a
2­
week
oral
exposure
study
conducted
by
NTP
(
1991),
depletion
(
atrophy)
and
necrosis
of
lymphoid
tissues
of
the
spleen,
thymus,
and
lymph
nodes
was
observed
in
B6C3F
1
mice
administered
lethal
concentrations
of
1,000
and
3,000
ppm
of
HCBD.
However,
the
investigators
noted
that
these
lesions
may
have
been
secondary
to
chemical­
induced
stress.
Similar
lesions
were
not
observed
in
mice
administered
19.2
mg/
kg­
day
in
a
subsequent
13­
week
study
conducted
by
NTP
(
1991).

In
a
13­
week
gavage
study,
relative
spleen
weights
were
significantly
increased
in
male
rats
orally
administered
HCBD
at
15.6
mg/
kg­
day
and
in
females
at
6.3
mg/
kg­
day
and
above
(
Harleman
and
Seinen,
1979).
Treatment­
related
lesions
in
lymphoid
organs
(
thymus,
lymph
nodes,
spleen)
have
not
been
reported
in
terminal
necropsy
of
mice
or
rats
in
other
HCBD
subchronic
and
chronic
oral
exposure
studies
at
doses
up
to
100
mg/
kg­
day
(
Harleman
and
Seinen,
1979;
Kociba
et
al.
1977;
Schwetz
et
al.,
1977).
Immune
function
screening
batteries
in
HCBD­
treated
animals
has
not
been
evaluated.

Delayed
hypersensitivity
reaction
was
exhibited
in
guinea
pigs
to
dermal
HCBD
application
(
Gradiski
et
al.,
1975).
HCBD
 
February
2003
7­
23
7.3.3
Hormonal
Disruption
No
studies
were
identified
that
associate
HCBD
exposure
with
endocrine
disruption.

7.3.4
Physiological
or
Mechanistic
Studies
Many
studies
have
been
undertaken
to
investigate
the
mechanisms
underlying
the
effects
of
HCBD
and,
in
particular,
its
toxicity
in
the
proximal
renal
tubule.
The
mode
of
action
for
HCBD
as
determined
from
these
studies
and
implications
for
cancer
assessment
are
discussed
in
Section
7.4.3
The
proximal
tubule­
specific
toxicity
of
HCBD
is
likely
determined
by
two
factors:
1)
the
distribution
of
enzymes
required
for
its
bioactivation,
and
2)
the
ability
of
this
region
to
concentrate
precursors
of
the
ultimate
toxic
species
(
Dekant
et
al.,
1990).
The
enzyme
cysteine
conjugate
$­
lyase
is
believed
to
catalyze
the
conversion
of
HCBD­
cysteine
conjugates
to
a
highly
reactive
thioketene
metabolite
(
Figure
6­
1,
Section
6.3).
Multiple
investigators
have
addressed
the
localization
of
$­
lyase
and
its
relationship
to
nephrotoxicity.
MacFarlane
et
al.
(
1989)
demonstrated
by
immunocytochemical
technique
that
the
region
of
highest
cytosolic
$­
lyase
activity
in
untreated
rats
coincides
with
the
site
of
HCBD­
induced
necrosis
in
the
pars
recta
region
of
the
proximal
tubule.
However,
Jones
et
al.
(
1988)
and
Kim
et
al.
(
1997)
detected
$­
lyase
in
the
entire
proximal
tubule.
Trevisan
et
al.
(
1998)
detected
histopathological
changes
and
increased
levels
of
$­
lyase
activity
in
the
urine
following
treatment
of
rats
with
S
3
and
S
1­
S
2
specific
nephrotoxicants,
which
were
cited
as
evidence
for
distribution
of
the
enzyme
along
the
entire
length
of
the
proximal
tubule.
These
data
suggest
that
additional
factors
may
contribute
to
selective
damage
in
the
pars
recta.

The
ability
of
the
proximal
tubule
to
concentrate
HCBD
metabolites
has
been
investigated
as
a
factor
in
renal
toxicity.
Nash
et
al.
(
1984)
administered
a
single
dose
of
radiolabeled
HCBD
and
observed
that
radioactivity
was
concentrated
in
the
renal
cortex
shortly
after
dosing.
Renal
cells
that
concentrated
the
radiolabeled
compounds
were
subsequently
observed
to
undergo
necrosis.
In
mammals,
(­
glutamyltranspeptidase,
the
enzyme
that
together
with
dipeptidase
catalyzes
the
conversion
of
the
glutathione
conjugate
to
the
cysteine
conjugate,
is
concentrated
in
the
brush­
border
membrane
of
the
proximate
tubular
cells.
The
distribution
of
this
enzyme
may
also
contribute
to
an
increase
in
the
concentration
of
cysteine
conjugates
in
the
proximal
renal
tubules.

The
probenecid­
sensitive
organic
anion
transporter
that
is
present
on
the
basolateral
side
of
proximal
tubule
cells
appears
to
play
a
role
in
the
accumulation
of
HCBD
metabolites
(
Dekant
1996).
Probenecid
is
a
competitive
inhibitor
of
organic
anion
transport,
and
is
reported
to
be
without
effect
on
energy
metabolism,
transport
carrier
synthesis,
or
uptake
of
other
substances
actively
transported
by
the
kidney
(
Lock
and
Ishmael,
1985;
Dekant,
1996).
Haloalkene­
derived
mercapturates
have
the
highest
affinity
for
the
organic
anion
transporter,
but
glutathione
and
cysteine
S­
conjugates
with
lipophilic
substituents
on
sulfur
are
also
substrates
for
transport
(
Dekant,
1996).

The
effect
of
probenecid
on
development
of
HCBD­
induced
renal
toxicity
has
been
investigated
in
in
vivo
and
in
vitro
studies.
Lock
and
Ishmael
(
1985)
administered
a
single
intraperitoneal
dose
of
16
or
64
:
mol/
kg
14C­
radiolabeled
N­
acetyl­
pentachlorobutadienyl­
L­
cysteine
to
female
Alpk/
AP
rats
and
observed
acute
renal
necrosis
within
four
hours.
Prior
administration
of
HCBD
 
February
2003
7­
24
up
to
500
:
mol/
kg
probenecid
reduced
renal
cortical
concentrations
of
radioactivity
and
provided
protection
against
nephrotoxicity
in
a
dose­
dependent
manner
as
assessed
by
plasma
urea
concentration
and
renal
histopathology.
Pretreatment
with
probenecid
also
reduced
or
prevented
the
toxic
effects
of
intraperitoneally
injected
HCBD
and
its
glutathione
and
cysteine
conjugates.
Thus,
the
selective
toxicity
to
the
pars
recta
in
rats
is
thought
to
result
in
part
from
transport
of
HCBD
metabolites
into
cells
of
this
region
via
a
probenecid­
sensitive
transport
system.

Bach
et
al.
(
1986)
confirmed
antagonism
of
probenecid
in
HCBD
metabolite­
induced
toxicity
in
freshly
isolated
rat
proximal
tubule
fragments.
Incorporation
of
3H­
proline
into
acid
precipitable
protein
was
utilized
as
an
indicator
of
synthetic
capacity
of
the
tubular
fragment.
Addition
of
2
mM
N­
acetyl­
pentachlorobutadienyl­
L­
cysteine
to
the
incubation
medium
reduced
3H­
proline
incorporation
to
34%
of
the
control
value.
The
inclusion
of
400
:
M
probenecid
in
the
incubation
medium
almost
completely
restored
(
to
95%)
3H­
proline
incorporation.

Multiple
studies
suggest
that
renal
cortical
mitochondria
are
a
primary
subcellular
target
for
HCBD
toxicity.
Jones
et
al.
(
1986)
investigated
the
toxic
effects
of
pentachlorobutadienylglutathione
(
PCBG)
in
isolated
rat
renal
epithelial
cells.
Exposure
to
PCBG
decreased
cell
viability
and
reduced
the
concentration
of
intracellular
thiols.
Other
PCBG­
related
effects
included
depletion
of
Ca2+
from
the
mitochondrial
compartment,
an
elevation
of
cytosolic
Ca2+
concentration,
inhibition
of
respiration,
and
decreased
levels
of
ATP.
Prevention
of
PCBG
bioactivation
by
inhibition
of
(­
glutamyl
transpeptidase
or
$­
lyase
provided
complete
protection
against
cytotoxicity.
The
authors
hypothesized
that
PCBG­
induced
renal
cell
injury
results
from
selective
effects
on
mitochondrial
function,
including
inhibition
of
respiration,
depression
of
ATP
synthesis,
and
release
of
mitochondrial
calcium
(
II)
ions.

Wallin
et
al.
(
1987)
studied
S­
pentachlorobutadienyl­
L­
cysteine
(
PCBC)
toxicity
in
mitochondria
isolated
from
the
rat
kidney
cortex.
Respiring
mitochondria
exposed
to
PCBC
showed
a
dose­
dependent
loss
of
ability
to
retain
calcium.
This
effect
was
associated
with
a
collapse
of
mitochondrial
membrane
potential.
A
slow
nonenzymatic
depletion
of
mitochondrial
glutathione
was
also
observed.
Preincubation
with
aminooxyacetic
acid,
an
inhibitor
of
$­
lyase,
effectively
counteracted
the
loss
of
glutathione,
suggesting
that
an
interaction
of
the
reactive
thioketene
with
the
mitochondrial
inner
membrane
was
responsible
for
the
observed
effects.

Schnellmann
et
al.
(
1987)
investigated
the
mechanism
of
PCBC­
induced
toxicity
in
renal
proximal
tubules
isolated
from
New
Zealand
rabbits.
Suspensions
of
isolated
tubules
were
exposed
to
concentrations
of
20
to
500
:
M
PCBC
in
the
presence
or
absence
of
chemicals
which
targeted
the
activity
of
ion­
channel
ATPases
(
nystatin,
ouabain),
Cytochrome
c
(
ascorbate,
tetramethylphenylenediamine),
cytochrome
oxidase
(
carbonyl
cyanide
ptrifluoromethoxyphenylhydrazone
and/
or
ATP
synthase
(
atractyloside
or
oligomycin).
Fifteen
minutes
of
exposure
to
200
:
M
PCBC
caused
an
increase
in
basal
and
ouabain­
insensitive
respiration
but
not
nystatin­
treated
respiration.
However,
sixty
minutes
of
PCBC
exposure
inhibited
basal,
nystatin­
stimulated
and
ouabain
insensitive
respiration,
and
resulted
in
a
79%
decrease
in
glutathione
concentration.
In
addition,
at
60
minutes,
an
11%
decrease
in
lactate
dehydrogenase
retention
was
observed,
suggesting
that
cell
viability
was
decreased
over
time
as
a
result
of
treatment.
The
changes
in
respiration
observed
at
60
minutes
appeared
to
result
from
gross
HCBD
 
February
2003
7­
25
mitochondrial
damage
characterized
by
inhibition
of
state
3
(
aerobic)
respiration,
depression
of
cytochrome
c/
cytochrome
oxidase
activity,
and
inhibition
of
electron
transport.
The
results
of
these
studies
suggest
that
alterations
in
mitochondrial
function
are
an
early
event
in
PCBC­
mediated
toxicity.

A
similar
pattern
of
events
was
observed
by
Groves
et
al.
(
1991).
These
workers
investigated
the
relationship
between
uptake
and
covalent
binding
of
the
HCBD
metabolite
pentachlorobutadienyl­
L­
cysteine
(
PCBC)
in
rabbit
renal
proximal
tubules,
renal
membrane
vesicles,
and
isolated
renal
cortical
mitochondria.
Their
findings
confirmed
the
PCBC­
induced
pattern
of
mitochondrial
dysfunction
previously
observed
by
Schnellmann
et
al.
(
1987)
in
rabbit
proximal
tubule
suspensions.
Furthermore,
Groves
et
al.
(
1991)
demonstrated
the
rapid
accumulation
of
35SPCBC
in
renal
proximal
tubule
cells
and
its
metabolism
to
a
reactive
intermediate
that
bound
to
tubular
protein.
An
estimated
70
to
90%
of
the
intracellular
radioactivity
was
bound
to
protein.
Mitochondria
isolated
from
renal
proximal
tubules
also
metabolized
35S­
PCBC
to
a
reactive
intermediate
that
bound
to
mitochondrial
protein,
consistent
with
the
mitochondrion
being
a
critical
subcellular
target
for
HCBD­
induced
toxicity.
Addition
of
the
$­
lyase
inhibitor
aminooxyacetic
acid
(
AOAA)
reduced
covalent
binding
to
tubular
proteins,
and
blocked
the
toxic
effects
of
PCBC
on
isolated
mitochondria.
However,
AOAA
decreased
but
did
not
prevent
the
toxic
effects
PCBC
on
respiration
and
cellular
ATP
levels
induced
by
PCBC
exposure.

Additional
studies
have
investigated
interactions
between
the
reactive
intermediate
generated
by
metabolism
of
PCBC
and
cellular
macromolecules.
Lock
and
Schnellmann
(
1990)
examined
the
ability
of
reactive
thiols
formed
by
the
action
of
$­
lyase
on
cysteine
conjugates
of
several
haloalkenes,
including
HCBD,
to
inhibit
renal
enzymes.
The
activities
of
glutathione
reductase
(
a
cytosolic
enzyme)
and
lipoyl
dehydrogenase
(
a
mitochondrial
enzyme)
were
assayed
for
this
purpose.
Administration
of
200
mg/
kg
HCBD
to
male
rats
by
intraperitoneal
injection
resulted
in
inhibition
of
both
enzymes.
The
authors
suggested
that
such
inhibition
is
a
general
outcome
of
PCBC
exposure,
and
is
likely
to
occur
with
a
diverse
range
of
renal
enzymes.

Schrenk
and
Dekant
(
1989)
investigated
covalent
binding
of
14C­
labeled
HCBD
metabolites
to
mouse
DNA
after
a
single
oral
dose
of
30
mg/
kg
14C­
HCBD.
HCBD
metabolites
bound
extensively
to
mitochondrial
DNA.
In
contrast,
little
binding
to
nuclear
DNA
was
observed.
The
study
authors
suggested
that
proximity
to
high
$­
lyase
concentration
in
the
mitochondrial
membrane
and
the
absence
of
associated
histones
make
mitochondrial
DNA
a
more
vulnerable
target
for
reactive
HCBD
metabolites.

As
noted
above
(
Section
6.3),
the
cysteine
derivative
of
HCBD
is
a
substrate
for
cysteine
conjugate
$­
lyase.
The
activity
of
$­
lyase
leads
to
formation
of
an
enethiol
intermediate
which
is
rapidly
converted
to
thioketene,
a
potent
acylating
agent
(
Dekant
et
al.,
1990).
In
rats,
the
enethiol
intermediate
may
be
detoxified
by
methylation
to
form
pentachlorobutadienyl­
methylthioether.
Morel
et
al.
(
1999)
investigated
the
role
of
S­
adenosyl
methionine
(
SAM)­
dependent
thiol
methylation
in
prevention
of
HCBD­
induced
nephrotoxicity
in
male
Swiss
OF1
mice.
The
mice
were
treated
with
a
single
intraperitoneal
dose
of
periodate­
oxidized
adenosine
(
ADOX)
prior
to
administration
of
a
single
intraperitoneal
dose
of
80
or
100
mg
HCBD/
kg.
Pretreatment
with
ADOX
increased
the
level
of
SAM
in
the
liver
and
kidney
approximately
four­
fold,
but
did
not
modify
the
HCBD
 
February
2003
7­
26
nephrotoxicity
of
HCBD
as
determined
by
histopathological
evaluation
of
renal
proximal
tubules.
This
result
was
interpreted
by
the
study
authors
as
evidence
that
SAM­
dependent
thiol
methylation
does
not
play
a
role
in
detoxification
of
HCBD­
derived
enethiol
in
mice.

Chemically­
induced
"
2:­
globulin
nephropathy
represents
a
potential
alternative
mechanism
for
HCBD
toxicity
in
rats.
Since
"
2:­
globulin
synthesis
is
androgen­
dependent
in
the
liver,
this
form
of
nephropathy
occurs
exclusively
in
male
rats,
and
is
characterized
by
the
accumulation
of
hyaline
droplets
in
proximal
tubule
cells.
Binding
of
the
chemical
to
"
2:­
globulin
is
a
prerequisite
for
development
of
nephrotoxicity.
Because
HCBD­
induced
renal
toxicity
occurs
in
both
male
and
female
rats,
it
is
evident
that
"
2:­
globulin
nephropathy
is
not
required
for
nephrotoxicity.
However,
there
is
limited
evidence
to
suggest
that
the
"
2:­
globulin
mechanism
may
contribute
to
HCBDinduced
nephrotoxicity
observed
in
male
rats.
Birner
et
al.
(
1995)
observed
that
unmetabolized
HCBD
was
excreted
in
the
urine
of
male,
but
not
female,
Wistar
rats
following
exposure
to
a
single
gavage
dose
of
14C­
HCBD
in
corn
oil.
The
study
authors
also
noted
more
pronounced
necrotic
changes
in
the
proximal
tubules
of
male
rats
when
examined
48
hours
after
treatment.
Slight
liver
damage
was
observed
only
in
male
rats.
In
a
subsequent
experiment
in
the
same
laboratory,
Pähler
et
al.
(
1997)
orally
administered
200
mg/
kg
14C­
HCBD
in
corn
oil
to
Sprague­
Dawley
(
SD)
and
NCI
Black­
Reiter
rats
(
NBR),
an
"
2:­
globulin­
deficient
strain..
14C­
HCBD
was
present
only
in
the
urine
of
male
SD
rats,
but
not
NBR
rats.
The
study
authors
determined
that
the
excreted
HCBD
detected
in
the
urine
of
male
SD
rats
was
associated
with
its
binding
to
"
2:­
globulin.
Histopathological
examination
48
hours
after
treatment
revealed
the
formulation
of
hyaline
droplets
indicative
of
"
2:­
globulin
accumulation
in
renal
epithelial
cells
of
male
SD
rats,
but
not
in
male
NBR
rats.
In
addition,
microscopic
examination
confirmed
the
occurrence
of
more
extensive
nephropathy
in
male
than
in
female
animals
as
previously
observed
in
Wistar
rats.

Saito
et
al.
(
1996)
established
that
dose­
dependent
levels
of
kidney­
type
"
2:­
globulin
("
G­
K)
in
the
urine
are
a
reliable
predictor
of
"
2:­
globulin
accumulation
in
the
kidney.
These
investigators
subsequently
administered
100
mg/
kg­
day
HCBD
to
adult
male
Sprague­
Dawley
rats
for
five
consecutive
days.
No
increase
in
urinary
"
G­
K
was
detected
following
exposure,
suggesting
that
HCBD
treatment
did
not
induce
a
marked
accumulation
of
"
2:­
globulin.
However,
histopathological
examination
revealed
some
epithelial
cells
showing
hyaline
droplet­
related
degeneration.
The
size
of
the
hyaline
droplets
formed
following
HCBD­
treatment
were
generally
smaller
than
those
observed
after
treatment
with
the
well­
characterized
"
2:­
globulin
nephropathy­
inducing
agent
dlimonene

No
significant
increase
in
"
2:­
globulin
was
observed
in
kidney
cytosol
prepared
from
Fischer
344/
N
rats
treated
with
a
single
oral
dose
of
200
mg
HCBD/
kg
when
assayed
by
Western
blot
and
capillary
electrophoresis
(
Pähler
et
al.,
1999).

"
2:­
globulin
accumulation
is
unique
to
male
rats,
and
when
in
excess
is
associated
with
nephropathy,
cytotoxicity,
cellular
proliferation,
and
consequent
renal
tumors.
HCBD
does
not
cause
an
accumulation
of
"
2:­
globulin
in
the
male
rats
(
Saito
et
al.,
1996;
Pähler
et
al.,
1999).
Moreover,
HCBD
clearly
causes
kidney
damage
in
both
male
and
female
rats
and
in
mice
(
Kociba
et
al.,
1977;
NTP,
1991).
With
respect
to
kidney
tumors,
Kociba
(
1977)
shows
that
kidney
tumors
are
found
in
both
HCBD­
treated
male
and
female
rats.
Based
upon
current
EPA
guidelines
(
U.
S.
EPA,
1991e),
HCBD
 
February
2003
7­
27
"
2:­
globulin
accumulation
does
not
appear
to
be
a
mode
of
action
of
HCBD­
induced
kidney
toxicity
in
rats
or
relevant
to
human
risk.

7.3.5
Structure­
Activity
Relationship
HCBD
appears
to
share
a
common
target
tissue
with
several
structurally­
related
haloalkenes,
including
perfluoropropene,
trichloroethene,
tetrachloroethene,
and
trichlorotrifluoropropene.
All
of
these
chemicals
demonstrate
dose­
dependent
toxicity
in
the
proximal
tubule.
Trichloroethene,
tetrachloroethene,
tetrafluoroethylene,
and
HCBD
have
all
been
found
to
induce
neoplasia
of
the
proximal
tubule
in
rats.
The
common
basis
for
toxicity
may
be
bioactivation
of
these
compounds
by
a
multistep
pathway
which
is
initiated
by
conjugation
with
glutathione,
resulting
in
the
formation
of
a
glutathione
S­
conjugate.
Metabolism
to
the
corresponding
cysteine
S­
conjugates,
and
subsequent
degradation
by
renal
cysteine
conjugate
$­
lyase,
yields
reactive
electrophiles
that
are
believed
to
be
ultimately
responsible
for
renal
toxicity.
These
electrophiles
alkylate
mitochondrial
macromolecules,
resulting
in
cellular
energy
deficit,
loss
of
membrane
potential,
and
disruption
of
calcium
homeostasis.

When
haloalkenes
are
considered
as
a
group,
the
extent
of
conjugation
is
much
higher
with
liver
microsomes
than
with
liver
cytosol,
in
contrast
to
results
observed
with
most
other
substrates.
This
effect
is
attributed
to
the
preferential
distribution
of
the
highly
lipophilic
haloalkenes
into
lipid
membranes,
thus
providing
high
substrate
concentrations
for
membrane­
bound
glutathione
Stransferase
(
Dekant
et
al.,
1990).
In
vitro
studies
suggest
that
rates
of
haloalkene
conjugation
correlate
well
with
the
chemical
reactivity
of
the
individual
compounds
(
Dekant
et
al.,
1990).
For
example,
substitution
with
chlorine
results
in
stabilization
of
a
B­
bond.
Chloroalkenes
are
thus
reported
to
be
more
resistant
to
metabolism
by
glutathione
conjugation
than
are
fluoroalkenes.

Investigations
have
compared
the
toxicity
of
structurally­
related
haloalkene
conjugates.
Lock
and
Schnellmann
(
1990)
investigated
the
effect
of
HCBD
and
other
haloalkene
cysteine
conjugates
on
renal
glutathione
reductase
and
lipoyl
dehydrogenase
activity,
and
concluded
that
inhibition
of
these
enzymes
by
the
reactive
thiols
formed
by
$­
lyase
cleavage
of
haloalkene
cysteine
conjugates
represented
a
general
mechanism
of
toxicity.

Green
and
Odum
(
1985)
investigated
the
nephrotoxicity
and
mutagenicity
of
the
cysteine
conjugates
of
halogenated
alkenes
in
rat
kidney
slices.
Compounds
investigated
included
the
chloroalkenes
HCBD,
trichloroethylene
and
perchloroethylene,
and
the
fluoroalkenes
hexafluoropropene
(
HFP)
and
tetrafluoroethylene
(
TFE).
All
of
these
conjugates
had
a
marked
effect
on
the
uptake
of
both
the
organic
anion
p­
aminohippuric
acid
(
PAH)
and
the
cation
tetraethylammonium
bromide
(
TEA)
into
rat
kidney
slices.
This
observation
was
considered
to
be
consistent
with
the
known
nephrotoxicity
of
HCBD,
TFE
and
HFP
in
vivo.
Each
of
the
conjugates
was
metabolized
by
rat
kidney
slices
and
by
semi­
purified
rat
kidney
$­
lyase
to
pyruvate,
ammonia,
and
an
unidentified
reactive
metabolite.
Although
all
of
the
conjugates
were
activated
by
$­
lyase
and
had
a
similar
effect
on
ion
transport,
their
mutagenicity
differed.
The
conjugates
of
HCBD,
trichloroethylene
and
perchloroethylene
were
mutagenic
in
the
Ames
bacterial
mutation
assay
when
activated
by
rat
kidney
S9
fraction.
In
contrast,
the
conjugates
of
TFE
and
HFP
were
not
mutagenic
either
in
the
presence
or
absence
of
rat
kidney
S9
fraction.
HCBD
 
February
2003
7­
28
Birner
et
al.
(
1997)
compared
the
nephrotoxicity
of
cysteine
S­
conjugates
derived
from
trichloroethene,
tetrachloroethene,
and
HCBD.
Male
and
female
rats
received
identical
intravenous
doses
of
S­(
1,2­
dichlorovinyl)­
L­
cysteine
(
1,2­
DCVC),
S­(
2,2­
dichlorovinyl)­
L­
cysteine
(
2,2­
DCVC),
S­(
1,2,2­
trichlorovinyl)­
L­
cysteine
(
TCVC),
or
S­(
1,2,3,4,4­
pentachlorobutadienyl)­
Lcysteine
(
PCBC).
Assessment
of
the
relative
nephrotoxic
potency
of
the
conjugates
by
histopathological
examination
and
excretion
of
(­
glutamyltranspepidases
in
urine
indicated
a
decrease
in
the
order
TCVC
>
1,2­
DCVC
>
PCBC
$
2,2­
DCVC.

7.4
Hazard
Characterization
7.4.1
Synthesis
and
Evaluation
of
Major
Noncancer
Effects
There
are
no
unconfounded
reports
of
human
health
effects
following
HCBD
exposure
by
any
route.
Oral
exposure
studies
of
HCBD
toxicity
in
animals
are
summarized
in
Table
7­
5.
A
distinctive
feature
of
HCBD
toxicity
in
animals
is
its
selective
effect
on
the
kidney,
regardless
of
the
route
of
administration.
Toxicity
within
the
kidney
is
also
selective,
with
damage
restricted
to
the
proximal
tubule.
In
rats,
damage
is
further
localized
to
the
pars
recta
region
of
the
proximal
tubule.

Subchronic
and
chronic
studies
in
rodents
present
a
clear
picture
of
dose­
related
renal
damage.
Progressive
events
over
time
include
changes
in
kidney
weight,
renal
tubular
degeneration,
necrosis
and
regeneration,
hyperplasia,
focal
adenomatous
proliferation,
and
tumor
formation.
Tumor
formation
occurs
exclusively
in
the
kidney,
and
only
at
doses
that
cause
extensive
cytotoxicity.

Evidence
from
metabolic
enzyme
inhibitor
studies,
cannulation
experiments,
and
analysis
of
urinary
metabolites
indicates
that
the
nephrotoxicity
of
HCBD
is
dependent
on
a
multistep
bioactivation
mechanism
involving
both
liver
and
kidney
enzymes.
The
initial
step
in
HCBD
metabolism
is
the
glutathione­
S­
transferase
mediated
biosynthesis
of
a
glutathione
conjugate
(
PCBG)
in
the
liver.
After
elimination
into
the
bile,
PCBG
undergoes
subsequent
metabolism
to
a
cysteine
conjugate
(
PCBC)
in
the
bile,
gut
or
kidneys.
PCBC
may
be
acetylated
by
renal
N­
acetyltransferases
to
form
a
N­
acetyl
cysteine
conjugates
(
N­
AcPCBC).
Both
PCBC
and
N­
AcPCBC
are
concentrated
in
renal
cells
via
an
active
transport
system
(
Dekant,
1990).
N­
AcPCBC
can
be
excreted
in
the
urine
or
de­
acetylated
to
regenerate
PCBC.
PCBC
is
a
substrate
for
$­
lyase­
dependent
activation
to
a
highly
reactive
thioketene
in
the
kidney.
Covalent
binding
of
this
reactive
species
to
cellular
macromolecules
is
believed
to
initiate
the
damage
that
ultimately
results
in
renal
cell
toxicity.

Potential
molecular
targets
for
binding
of
the
reactive
thioketene
include
enzymes,
membrane
proteins,
glutathione,
phospholipids,
and
mitochondrial
DNA.
Localized
damage
to
the
proximal
tubule
is
believed
to
reflect
high
$­
lyase
concentration
in
this
region.
Evidence
from
studies
using
the
selective
inhibitor
probenecid
suggests
that
accumulation
of
the
cysteine
and
N­
acetyl
cysteine
conjugates
via
anion
transport
systems
localized
in
this
segment
of
the
proximal
tubule
may
account
for
this
selective
pattern
of
toxicity.

In
vitro
studies
suggest
that
cortical
mitochondria
are
the
critical
subcellular
target
for
toxicity
of
the
bioactivated
sulfur
conjugates
of
HCBD.
Susceptibility
of
mitochondria
to
HCBD
toxicity
is
linked
to
high
concentrations
of
$­
lyase
associated
with
mitochondrial
membranes.
The
HCBD
 
February
2003
7­
29
Table
7­
5.
Summary
of
Principal
HCBD
Toxicity
Studies.

Species/
Strain
Number/

Sex/
Dose
Route
Frequency/

Duration
Doses
(
mg/
kg­
day)
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
Effect(
s)
Reference
ACUTE
EXPOSURE
Rat
6
M
G
Single
Dose
24
hours
200
­­
200
Increased
plasma
urea;

increased
proteins
and
metabolites
in
urine
Nash
et
al.

(
1984)

Rat
/
Wistar
5
M
G
Single
Dose
24
hours
0,
10,
100,

200
10
100
Limited
focal
necrosis
of
the
kidneys
and
other
signs
of
renal
damage
in
urine
and
blood
Jonker
et
al.

(
1993a)

Rat
/
Sprague­

Dawley
4­
5
M
G
Single
Dose
24
hours
0,
100,
200
­­
100
Histological
kidney
lesions
and
impaired
kidney
function
Payan
et
al.

(
1993)

SHORT­
TERM
EXPOSURE
Rat
/
Sprague­

Dawley
4
F
F
Daily
30
days
0,
1,
3,
10,
30,

65,
100
3
10
(
reduced
body
wt.

gain;
increased
hemoglobin
concentration)
Increased
relative
kidney
weight;
renal
tubular
degeneration,
necrosis
and
regeneration;
increased
hemoglobin
concentration
reduced
body
weight
gain
Kociba
et
al.

(
1971)

Rat
/
Wistarderived
6
M
6
F
F
Daily
14
days
0,
4.6,
14,

35.3
­­
4.6
Degeneration
of
renal
tubular
epithelial
cells,

localized
to
pars
recta
Harleman
and
Seinen
(
1979)

Rat
/
Sprague­

Dawley
5
M
G
Daily
3
weeks
0,
0.2,
20
0.2
20
Histopathological
indications
of
renal
cortical
damage;
decreased
body
weight
gain;
increased
relative
kidney
weight
Stott
et
al.

(
1981)
Table
7­
5
(
continued)

Species/
Strain
Number/

Sex/
Dose
Route
Frequency/

Duration
Doses
(
mg/
kg­
day)
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
Effect(
s)
Reference
HCBD
 
February
2003
7­
30
Mouse
/

B6C3F
1
5
M
5
F
F
Daily
15
days
M:
0,
3,
12,
40
F:
0,
5,
16,
49
­
3
5
Renal
necrosis
and
cellular
regeneration
NTP
(
1991)

Yang
et
al.

(
1989)

Rat
/
Wistar
5
M
5
F
F
Daily
4
weeks
0,
2.25,
8,
28
2.25
8
Renal
tubular
cytomegaly;

decreased
plasma
creatine;

decreased
body
weight,

decreased
liver
and
adrenal
weight
Jonker
et
al.

(
1993b)

Rat
/
Wistar
12
M
F
Daily
3
weeks
0,
7.2,
36,
180
7.2
36
(
body
weight)

180
(
kidney
lesions)
Lower
mean
body
weight
(
15%
decrease);
extensive
regeneration
at
180
but
not
36
mg/
kg­
day
Nakagawa
et
al.
(
1998)

SUBCHRONIC
EXPOSURE
Rat
/
Wistarderived
10
M
10
F
G
Daily
13
weeks
0,
0.4,
1.0,

2.5,
6.3,
15.6
2.5
(
M)

1.0
(
F)
6.3
(
M)

2.5
(
F)
Proximal
tubular
degeneration
Harleman
and
Seinen
(
1979)

Mouse
/

B6C3F
1
10
M
10
F
F
Daily
13
weeks
M:
0,
0.1,
0.4,

1.5,
4.9,
16.8
F:
0,
0.2,
0.5,

1.8,
4.5,
19.2
1.5
(
M)

0.2?
(
F)
4.9
(
M)

0.5
(
F)
Renal
tubular
regeneration
NTP
(
1991)
Table
7­
5
(
continued)

Species/
Strain
Number/

Sex/
Dose
Route
Frequency/

Duration
Doses
(
mg/
kg­
day)
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
Effect(
s)
Reference
HCBD
 
February
2003
7­
31
Rat
/
Sprague­

Dawley
Gross
Pathology:

10­
12
M,

20­
24
F
(
HCBD)
17
M,
34
F
(
control)
Histopatho
logy:

5
M
5
F
F
Daily
0.2,
2.0,
20
0.2
2
Renal
tubular
hyperplasia
Schwetz
et
al.
(
1977)

Rat
/
Wistar
21
M
F
Daily
30
weeks
90
­­
90
Decreased
final
weight
and
increased
relative
kidney
weight
Nakagawa
et
al.
(
1998)

CHRONIC
EXPOSURE
Rat
/
Sprague­

Dawley
39M,
40F
F
Daily
22
months
(
M)

24
months
(
F)
0,
0.2,
2,
20
0.2
2
Increased
urinary
coproporphyrin
excretion
(
females);
increased
kidney
weight;
renal
tubular
epithelial
hyperplasia
Kociba
et
al.

(
1977)
Table
7­
5
(
continued)

Species/
Strain
Number/

Sex/
Dose
Route
Frequency/

Duration
Doses
(
mg/
kg­
day)
NOAEL
(
mg/
kg­
day)
LOAEL
(
mg/
kg­
day)
Effect(
s)
Reference
HCBD
 
February
2003
7­
32
DEVELOPMENTAL
AND
REPRODUCTIVE
STUDIES
Rat
/
Sprague­

Dawley
10­
12
M,

20­
24
F
(
HCBD)

17M,
34
F
(
control)
F
Daily
148
days
(
90
days
prior
to
mating
through
postnatal
day
21)
0,
0.2,
2,
20
2
20
Slightly
decreased
neonatal
weight
Schwetz
et
al.
(
1977)

Rat
6
F
F
Daily
3
weeks
prior
to
mating;
3
weeks
during
mating;
throughout
gestation
and
lactation
0,
15,
150
­­
15
Maternal
toxicity
(
renal,

neurological).
Reduced
pup
weights
on
days
0,
10,
and
20
Harleman
and
Seinen
(
1979)

Rat
F
NS
NS
8.1
­­
8.1
Reduced
body
weight;

shorter
crown­
rump
length;

ultrastructural
chances
in
neurocytes;
increased
levels
of
free
radicals
in
brain
and
spinal
cord
Badaeva
(
1983)

Abbreviations:
Sex:
M
=
male
Route:
G
=
gavage
NS
=
Not
specified
F
=
female
F
=
feed
study
HCBD
 
February
2003
7­
33
reactive
metabolite
formed
by
$­
lyase
cleavage
of
sulfur
conjugates
is
thought
to
interact
with
components
of
the
inner
mitochondrial
membrane.
Disruption
of
respiration
and
uncoupling
of
oxidative
phosphorylation
leads
to
a
marked
reduction
of
ATP
levels
in
susceptible
kidney
cells,
and
ultimately
necrosis.

The
mechanism
described
above
is
believed
to
contribute
to
the
renal
damage
observed
in
both
male
and
female
rats.
However,
additional
mechanisms
may
contribute
to
nephrotoxicity
of
HCBD
in
male
rats.
Current
evidence
suggests
that
at
least
two
discrete
male­
specific
pathways
may
participate
in
the
more
pronounced
necrotic
changes
observed
in
the
renal
tubules
of
male
rats
in
some
studies.
Formation
of
hyaline
droplets
indicative
of
"
2:­
globulin
accumulation
has
been
observed
in
the
kidney
of
HCBD­
treated
male
rats.
The
significance
of
this
finding
for
HCBDinduced
nephrotoxicity
remains
to
be
determined.
A
second
potential
mechanism
for
male
specific
toxicity
involves
the
cytochrome
P450
3A­
mediated
formation
of
an
N­
acetylated
cysteine
conjugate
sulfoxide.

Other
noncancer
effects
associated
with
HCBD
exposure
in
animals
include
developmental
effects
and
neurotoxicity.
Reproductive
effects
were
observed
only
at
maternal
toxic
dose.
In
one
study,
female
Wistar
rats
were
administered
a
diet
containing
0,
15
or
150
mg/
kg­
day)
HCBD
for
3
weeks
prior
to
mating,
3
weeks
during
mating
and
throughout
gestation
and
lactation.
Maternal
toxicity
was
evident
in
treated
groups.
No
conceptions
occurred
for
the
high
dose
group,
the
ovaries
showed
little
follicular
activity,
and
no
uterine
implantation
sites
were
observed.
At
15
mg/
kg­
day,
pups
exhibited
lower
birth
weights
and
reduced
growth
compared
to
controls
(
Harleman
and
Seinen,
1979).
In
a
Russian
study
cited
through
secondary
sources,
pregnant
rats
administered
8.1
mg/
kg­
day
of
HCBD
during
gestation
gave
birth
to
pups
with
lower
body
weights
and
shorter
crown­
rump
lengths
(
Badaeva,
1983).

Harleman
and
Seinen
(
1979)
observed
ataxia,
incoordination,
weakness
of
the
hind
legs,
and
unsteady
gait
in
conjunction
with
demyelination
and
fragmentation
of
femoral
nerve
fibers
in
female
rats
consuming
dietary
dose
of
150
mg/
kg­
day
HCBD
for
10
to
18
weeks.
No
neurotoxic
effects
were
reported
for
rats
consuming
15
mg/
kg­
day.
In
a
Russian
study
cited
through
secondary
sources,
daily
oral
administration
of
8.1
mg/
kg­
day
HCBD
to
pregnant
rats
throughout
gestation
resulted
in
histopathological
changes
in
nerve
cells
and
myelin
fibers
of
the
spinal
cord
in
treated
dams
and
their
offspring
(
Badaeva
et
al.,
1985).

The
mode
of
neurotoxic
action
has
not
been
studied.
Toxicokinetic
studies
in
animals
following
oral
administration
demonstrated
that
HCBD
and
its
metabolites
distributed
to
the
brain
and
adipose
tissues
in
addition
to
the
kidney
and
the
liver
(
Reichert,
1983;
Reichert
et
al.,
1985;
Dekant
et
al.,
1988a).
Thus,
reported
toxicity
at
targets
other
than
the
kidney
may
be
related
to
the
distribution
of
HCBD
and/
or
its
metabolites
to
these
targets
and
subsequent
covalent
binding
of
the
reactive
metabolites
to
cellular
macromolecules.

An
important
issue
in
the
evaluation
of
the
hazard
posed
by
HCBD
concerns
the
applicability
of
mechanistic
data
obtained
in
rodent
studies
to
humans.
Limited
data
from
in
vitro
studies
with
human
renal
cytosol
and
cultured
human
proximal
tubule
cells
suggest
that
humans
have
the
ability
HCBD
 
February
2003
7­
34
to
form
HCBD
glutathione
conjugates
and
to
metabolize
HCBD
cysteine
conjugates
to
a
toxic
metabolite.

7.4.2
Synthesis
and
Evaluation
of
Carcinogenic
Effects
No
studies
of
the
potential
for
HCBD
carcinogenicity
in
humans
have
been
reported.
In
animals,
one
lifetime
exposure
carcinogenicity
study
has
been
performed.
Kociba
et
al.
(
1977)
observed
increased
incidence
of
renal
tumor
formation
in
male
and
female
rats
following
lifetime
exposure
to
HCBD
in
the
diet.
Neoplastic
changes
occurred
only
at
the
highest
dose,
which
exceeded
the
maximum
tolerated
dose
(
MTD).
There
was
increased
mortality,
significant
weight
loss
(
greater
than
10%),
and
severe
renal
toxicity.
This
pattern
suggests
that
tumor
formation
may
be
secondary
to
HCBD­
induced
cytotoxicity.
This
conclusion
is
supported
by
the
study
of
Nakagawa
et
al.
(
1998),
who
found
increased
cell
proliferation
and
increased
DNA
synthesis
in
the
outer
stripe
and
cortex
of
kidneys
from
HCBD­
treated
rats.

However,
available
data
must
be
considered
as
too
limited
to
support
a
conclusion
with
high
confidence.
The
widely­
spaced
doses
(
a
10­
fold
spacing
between
the
highest
and
next
lower
dose)
in
the
Kociba
et
al.
(
1977)
study,
for
example,
did
not
provide
the
opportunity
to
confirm
that
pronounced
cytotoxicity
is
a
prerequisite
for
tumorigenesis.
Additional
limitations
in
the
database
include
the
absence
of
cell
proliferation
studies
and
limited
in
vivo
data
for
mutagenesis.
These
limitations
prevent
the
use
of
cell­
kinetic
multistage
(
CKM)
models
in
the
analysis
of
cancer
risk
(
Bogen,
1989).

Results
from
mutagenicity
studies
with
HCBD
are
mixed.
In
the
presence
of
appropriate
metabolic
activation
conditions,
HCBD
and
its
metabolites
are
mutagenic
in
some,
but
not
all,
studies.
Thus,
a
genotoxic
mode
of
action
must
be
considered.
The
observation
that
high
doses
of
HCBD
metabolites
can
bind
to
DNA
weakly
in
vivo
in
mice
(
Schrenk
and
Dekant,
1989)
strengthens
this
conclusion
somewhat.
There
is,
however,
still
some
question
as
to
whether
genotoxicity
is
the
primary
mode
of
action
at
the
lower
concentrations
at
which
nephrotoxic
effects
are
first
observed.
The
weight
of
evidence
suggests
that
both
genotoxic
and
nongenotoxic
effects
contribute,
possibly
at
different
dose
levels,
and
that
the
current
database
is
insufficient
to
disentangle
the
relative
contributions
of
these
effects.

7.4.3
Mode
of
Action
and
Implications
in
Cancer
Assessment
Both
sustained
cytotoxic
damage
and
irreversible
DNA
binding
have
been
proposed
as
events
in
HCBD
renal
carcinogenesis
in
rodents
(
Stott
et
al.,
1981)
An.
A
lifetime
oral
study
in
rats
showed
kidney
tumors
at
a
very
high
dose
that
exceeded
the
MTD,
suggesting
that
HCBD­
induced
cytotoxicity
may
lead
to
tumor
formation
(
see
Section
7.4.2).
Studies
in
rats
and
mice
indicate
that
kidney
is
the
target
organ.
Progressive
toxicological
changes
are
observed
in
kidney
over
time:
decreased
and
increased
kidney
weight,
increased
excretion
of
coporphyrin
(
kidney
dysfunction),
renal
tubular
degeneration,
necrosis
and
regeneration,
hyperplasia,
focal
adenomatous
proliferation,
and
finally
tumor
formation.
HCBD
 
February
2003
7­
35
On
the
other
hand,
in
the
presence
of
metabolic
activation,
HCBD
and
its
reactive
metabolites
are
mutagenic
in
some
(
Simmon,
1977;
Reichert
et
al.,
1984;
Reichert
and
Schutz,
1986;
Wild
et
al.,
1986),
but
not
all,
studies
(
See
Section
7.3.1).
Those
studies
done
either
on
kidney
cells
(
in
vivo)
or
with
kidney
microsomes
(
in
vitro)
appear
to
represent
most
of
the
mutagenic
response
to
HCBD.
Thus,
a
mutagenic
mode
of
action
cannot
be
ruled
out
(
Dekant
et
al.,
1990;
Lock,
1994).

The
hypothesis
that
both
cytotoxicity
and
mutagenic
mode
of
action
may
be
operating
is
consistent
with
the
findings
that
the
adverse
effects
of
HCBD
are
dependent
on
a
multistep
pathway
of
bioactivation.
The
ultimate
step
in
this
pathway
is
a
$­
lyase­
mediated
degradation
of
a
HCBD
metabolite
that
generates
a
highly
reactive
thioketene
in
proximal
tubule
cells.
Covalent
binding
of
this
thioketene
to
DNA,
proteins
and
other
macromolecules
is
considered
to
be
the
mechanism
responsible
for
the
observed
cytotoxic
and
mutagenic
effects
of
HCBD
and
its
metabolites.
Restriction
of
these
effects
to
the
proximal
tubule
most
likely
reflects
both
uptake
processes
that
concentrate
the
cysteine
conjugate
substrate
in
epithelial
cells,
and
localization
of
(­
glutamyltranspeptidase
and
$­
lyase
activity
to
this
region
of
the
kidney.

In
vitro
studies
(
Schnellman
et
al.,
1987;
Groves
et
al.,
1991;
Jones
et
al.,
1986;
Wallin
et
al.,
1987)
indicate
that
mitochondria
in
renal
tubular
epithelial
cells
are
the
major
target
for
HCBD
metabolite­
induced
toxicity.
The
reactive
metabolite
formed
by
$­
lyase
cleavage
of
cysteine
conjugate
is
thought
to
interact
with
components
of
the
mitochondrial
inner
membrane.
The
initial
effect
is
an
uncoupling
of
oxidative
phosphorylation
and
prevention
of
generation
of
ATP.
The
decrease
in
renal
tubular
ATP
secondary
to
mitochondrial
dysfunction
in
turn
limits
ATP
dependent
active
transport
in
the
tubules,
inhibiting
reabsorption
processes
(
Jaffe
et
al.,
1983).
Later
effect
involves
gross
mitochondrial
damage
characterized
by
inhibition
of
cytochrome
c­
cytochrome
oxidase
activity,
and
inhibited
electron
transport.
This
sequence
of
events
in
the
renal
proximal
tubules
ultimately
leads
to
cell
death.
Other
studies
indicate
that
the
reactive
species
generated
by
$­
lyase­
mediated
degradation
of
HCBD
metabolites
interact
directly
with
mitochondrial
DNA
(
mtDNA)
from
mouse
kidney
(
Schrenk
and
Dekant,
1989).
Renal
mtDNA
may
be
the
preferential
target
due
to
the
high
concentration
of
$­
lyase
in
the
mitochondrial
membrane,
the
lack
of
protective
histones
associated
with
mitochondrial
DNA
(
Borst
&
Grivell,
1978),
and
an
inadequate
repair
function
(
Mansouri
et
al.,
1997).
Mutations
in
the
mtDNA
can
lead
to
a
respiratory
chain
deficiency
and
cell
dysfunction
when
the
percentage
of
the
mutants
reach
a
certain
level
(
Schapira,
1999).

Three
important
aspects
of
mitochondrial
oxidative
phosphorylation
involved
in
the
pathogenesis
of
mitochondrial
dysfunction
are:
generation
of
cellular
energy
in
the
form
of
ATP;
generation
of
reactive
oxygen
species
(
ROS);
and
regulation
of
apoptosis
or
programmed
cell
death
(
Wallace,
1999).
The
process
of
oxidative
phosphorylation
produces
significant
amounts
of
ROS
which
are
toxic
byproducts
of
respiration.
Chronic
exposure
to
ROS
can
result
in
oxidative
damage
to
mitochondrial
and
cellular
proteins,
and
mutations
in
the
mtDNA.
Because
mtDNA
codes
for
22
transfer
RNAs
(
tRNA)
and
2
ribosomal
RNAs
(
rRNA)
for
synthesis
of
important
mitochrondrial
proteins
involved
in
the
oxidative
phosphorylation,
functional
mtDNA
is
critical
to
the
normal
function
of
a
cell.
Mutations
in
mtDNA
may
lead
to
overexposure
to
ROS
and
decreased
energy
production.
Apoptosis
is
initiated
when
the
mitochondrial
permeability
transition
pore
(
mtPTP)
in
the
inner
membrane
opens
and
cell
death­
promoting
factors
such
as
the
caspases
are
released
(
Wallace,
1999).
Opening
of
the
mtPTP
and
the
accompanying
cell
death
can
be
initiated
by
the
HCBD
 
February
2003
7­
36
mitochondrion's
excessive
uptake
of
Ca2+,
increased
exposure
to
ROS,
or
decline
in
energetic
capacity.
Therefore,
a
marked
reduction
in
mitochondrial
energy
production
and
a
chronic
increase
in
oxidative
stress
could
activate
the
mtPTP
and
initiate
apoptosis.

Numerous
mtDNA
mutations
have
been
associated
with
human
mitochondrial
disease.
Mitochondrial
disease
is
a
disruption
of
the
proper
function
of
the
mitochondria,
resulting
in
a
variety
of
clinical
manifestation.
This
disruption
can
include
an
inhibition
of
the
electron
transport
chain,
a
disruption
of
oxidative
phosphorylation
and
an
increase
in
the
production
of
reactive
oxygen
species.
MtDNA
mutations
could
contribute
to
neoplastic
transformation
by
changing
cellular
energy
capacities,
mitochondrial
oxidative
stress,
and/
or
modulating
apoptosis
(
Wallace,
1999).
Thus,
it
may
be
postulated
that
mutations
of
renal
mtDNA
induced
by
HCBD
may
result
in
reduction
in
energy
production,
increase
in
oxidative
stress,
and
initiation
of
apotosis,
leading
to
tumor
formation.

Mitochondrial
dysfunction
may
also
result
from
interaction
of
highly
reactive
HCBD
metabolites
with
components
of
the
mitochondrial
inner
membrane,
such
as
enzymes
related
to
cell
function.
Subsequent
energy
depletion
may
trigger
the
renal
cytotoxicity
that
is
the
putative
mechanism
for
HCBD­
mediated
carcinogenesis.
Thus,
HCBD
induced
cytotoxicity
and
tumorigensis
may
be
ultimately
the
consequence
of
mitochondrial
dysfunction
resulting
from
exposure.

Recent
evidence
suggests
that
HCBD­
induced
"
2:­
globulin
accumulation
contributes
to
renal
injury
in
male
rats.
However,
renal
tubular
necrosis
and
renal
tubular
tumors
were
observed
in
both
male
and
female
rats
following
HCBD
exposure
(
Kociba
et
al.,
1977),
and
renal
necrosis
and
regeneration
were
also
observed
in
male
and
female
mice
(
NTP,
1991).
Therefore,
"
2:­
globulin
accumulation
cannot
be
the
sole
mechanism
for
HCBD­
induced
carcinogenesis.

7.4.4
Weight
of
Evidence
Evaluation
for
Carcinogenicity
No
human
carcinogenicity
data
are
available
for
HCBD.

A
single
lifetime
study
of
HCBD
carcinogenicity
in
rats
(
Kociba
et
al.,
1977)
is
available
for
evaluation.
This
study
revealed
statistically
significant
increases
in
the
incidence
of
tumors
in
male
and
female
rats
following
oral
HCBD
exposure.
Although
human
carcinogenicity
data
are
unavailable,
evidence
exists
that
the
metabolic
enzymes
responsible
for
conversion
of
HCBD
to
the
reactive
and
toxic
thioketene
occur
in
humans,
albeit
at
levels
lower
than
that
in
the
rat
(
see
Section
6.3).
In
accordance
with
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1986),
HCBD
is
best
classified
as
Group
C,
possible
human
carcinogens,
based
on
limited
evidence
of
carcinogenicity
in
one
animal
study,
and
no
data
in
humans.
Based
on
the
proposed
guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
19991996a),
HCBD
is
classified
as
"
likely
to
be
carcinogenic
to
humans
by
the
oral
route
of
exposure,
but
whose
carcinogenic
potential
by
the
inhalation
and
dermal
routes
of
exposure
cannot
be
determined
because
there
are
inadequate
data
to
perform
an
assessment".
This
descriptor
is
considered
appropriate
when
there
are
no
or
inadequate
data
in
humans,
but
the
combined
experimental
evidence
demonstrates
the
production
or
anticipated
production
of
tumors
in
animals
by
modes
of
action
that
are
relevant
or
assumed
to
be
relevant
to
humans,
tempered
by
the
lack
of
adequate
inhalation
or
dermal
studies.
Mechanistic
studies
performed
in
vitro
and
in
vivo
suggest
that
either
genotoxic
or
non­
genotoxic
modes
of
action
may
HCBD
 
February
2003
7­
37
underlie
or
contribute
to
the
carcinogenic
potential
of
HCBD.
However,
the
strongly
non­
linear
doseresponse
in
the
Kociba
et
al.
(
1977)
study
cannot
be
ignored.
At
present,
the
mode­
of­
action
information
still
lacks
identification
of
the
sequence
of
key
events
and
a
quantitative
description
of
the
doses
at
which
those
key
events
begin
to
occur.
In
such
cases,
EPA's
proposed
cancer
guideline
revisions
(
U.
S.
EPA,
1999)
support
consideration
of
both
linear
and
nonlinear
extrapolation
to
lower
doses.
In
the
absence
of
adequate
data
to
exclude
a
linear
mechanism(
s)
of
tumor
formation,
the
quantitative
cancer
risk
assessment
of
HCBD
should
conservatively
be
conducted
using
the
linear­
default
model.
Therefore,
both
linear
and
non­
linear
approaches
are
presented.

7.4.5
Sensitive
Populations
Sensitive
populations
are
those
which
experience
more
adverse
effects
at
comparable
levels
of
exposure,
or
which
experience
adverse
effects
at
lower
exposure
levels,
than
the
general
population.
The
enhanced
response
of
these
sensitive
subpopulations
may
result
from
intrinsic
or
extrinsic
factors.
Factors
that
may
be
important
include,
but
are
not
limited
to:
impaired
function
of
detoxification,
excretory,
or
compensatory
processes
that
protect
against
or
reduce
toxicity;
differences
in
physiological
protective
mechanisms;
genetic
differences
in
metabolism;
developmental
stage;
health
status;
gender;
or
age
of
the
individual.

Human
populations
that
exhibit
greater
sensitivity
to
HCBD
have
not
been
identified.
However,
it
has
been
generally
observed
that
existing
nephropathy
or
age­
related
kidney
degeneration
can
increase
the
risk
of
renal
injury
or
exacerbate
nephrotoxicity
in
humans
(
WHO,
1991).
Evidence
that
existing
nephropathy
increases
sensitivity
to
HCBD
toxicity
has
been
obtained
in
a
study
conducted
in
male
Wistar
rats
(
Kirby
and
Bach,
1995).
Nephrosis
was
induced
by
pretreatment
with
adriamycin
(
ADR),
and
rats
were
subsequently
exposed
to
HCBD.
Damage
to
the
proximal
tubule
was
more
severe
and
renal
cortical
repair
capacity
was
decreased
in
ADR­
treated
rats
when
compared
to
rats
exposed
to
HCBD
without
prior
ADR
exposure.
These
results
suggest
that
individuals
with
existing
kidney
damage
or
the
elderly
may
be
potentially
sensitive
populations
for
HCBD
exposure.

Studies
in
animals
showed
that
the
young
rats
and
mice
experience
acute
effects
at
significantly
lower
doses
than
do
adults
(
Hook
et
al.,
1983;
Lock
et
al.,
1984),
suggesting
that
infants
may
represent
a
potentially
sensitive
subpopulation
for
acute
HCBD
exposure,
perhaps
as
a
result
of
immature
organ
systems.
Additionally,
female
rodents
were
apparently
more
sensitive
than
males
to
acute
HCBD
exposure
(
Kociba,
1977).
The
mechanism
underlying
this
sensitivity
is
not
currently
clear.
2
Note
that
the
literature
has
used
the
terms
BMD
and
BMDL
in
a
confusing
way
(
Crump,
1984,1995).
The
EPA
benchmark
dose
software
(
BMDS
version
1.3.1)
and
EPA
technical
guidance
on
this
subject
(
U.
S.
EPA,
2000a),
use
the
term
"
BMD"
to
refer
to
the
central
or
maximum
likelihood
estimate
(
MLE)
of
the
dose
that
is
expected
to
yield
the
BMR.
"
BMC"
(
benchmark
concentration)
is
a
term
that
is
sometimes
used,
as
opposed
to
"
BMD,"
to
distinguish
between
inhalation
and
oral
benchmarks.
"
BMDL"
or
"
BMCL"
refer
to
the
lower
end
of
a
one­
sided
confidence
interval
for
that
central
estimate.
"
BMD"
will
be
used
to
refer
to
the
entire
process.
The
POD
for
low
dose
extrapolation
on
for
setting
the
RfD/
RfC
will
be
the
BMDL
or
BMCL.
To
simplify
further
discussion
in
this
document,
we
will
use
BMD
or
BMDL
generically
to
mean
oral
or
inhalation
values,
unless
stated
otherwise.

HCBD
 
February
2003
8­
1
8.0
DOSE­
RESPONSE
ASSESSMENT
8.1
Dose­
Response
for
Noncancer
Effects
8.1.1
RfD
Determination
The
reference
dose
(
RfD)
for
a
chemical
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).
Data
on
the
non­
cancer
effects
of
HCBD
from
chronic
and
subchronic
studies
were
used
to
estimate
a
RfD
value
using
the
benchmark
dose
(
BMD2)
approach
(
U.
S.
EPA,
1995).

Choice
of
Principal
Study
and
Critical
Effect
There
are
no
reliable
dose­
response
data
for
humans
exposed
to
HCBD.
The
NTP
(
1991)
subchronic
study
on
mice
was
chosen
as
the
principal
study,
with
the
Kociba
et
al.
(
1977)
and
Schwetz
et
al.
(
1977)
studies
on
rats
as
supporting
studies,
due
to
their
exposure
durations
and
sensitivity
of
endpoints
observed.
Hyperplasia
and
regeneration
of
the
renal
tubular
cell
epithelial
cells
was
selected
as
the
critical
effect,
based
upon
its
observation
at
the
lowest
doses.
The
RfD
for
HCBD
is
derived
from
a
BMDL
of
0.1
mg/
kg­
day
for
renal
tubular
epithelial
cell
hyperplasia/
regeneration
from
the
NTP
study.

In
the
Kociba
et
al.
(
1977)
lifetime
oral
exposure
study
of
rats
to
HCBD,
a
NOAEL
of
0.2
mg/
kg­
day
and
a
LOAEL
of
2
mg/
kg­
day
were
identified,
based
on
an
increase
in
renal
tubular
epithelial
cell
hyperplasia/
regeneration.
In
the
Schwetz
et
al.
(
1977)
148­
day
oral
exposure
study
of
rats
to
HCBD,
a
NOAEL
of
0.2
mg/
kg­
day
and
a
LOAEL
of
2
mg/
kg­
day
were
identified,
based
on
an
increase
in
the
severity
of
renal
tubular
dilation,
collapse,
and
atrophy.
In
the
13­
week
feeding
study
by
NTP
(
1991),
the
study
authors
identified
a
NOAEL
of
1.5
mg/
kg­
day
for
male
mice,
and
did
not
identify
a
NOAEL
for
female
mice
because
renal
tubular
regeneration
occurred
in
1
of
10
females
in
the
lowest
dose
group
(
0.2
mg/
kg­
day).
However,
others
(
U.
S.
EPA,
1998a;
WHO,
1994)
have
concluded
that
the
effect
observed
at
0.2
mg/
kg­
day
is
not
statistically
significant,
and
therefore
considered
this
dose
to
be
the
NOAEL.

In
order
to
avoid
the
uncertainty
in
the
LOAEL/
NOAEL
classification
of
the
0.2
mg/
kg­
day
dose
in
the
NTP
study,
the
most
sensitive
response
found,
a
benchmark
dose
(
BMD)
analysis
was
3
A
copy
of
BMDS
can
be
obtained
from
the
Internet
at
http://
www.
epa.
gov/
ncea/
bmds.
htm.

4
A
benchmark
response
(
BMR)
of
10%
was
used
for
all
BMD
analyses.

HCBD
 
February
2003
8­
2
conducted
on
the
NTP
female
mouse
renal
tubular
regeneration
response
(
Table
8­
1).
Using
BMDS3,
version
1.3.1,
the
data
was
fit
to
all
available
dichotomous
models
(
Table
8­
2).
The
data
was
best
fit
by
a
Weibull
model
(
p
=
1.00,
AIC
=
17),
resulting
in
a
BMD4
of
0.2
mg/
kg­
day
and
a
BMDL
of
0.1
mg/
kg­
day
(
Figure
8­
1).
This
dose
was
therefore
taken
as
the
point
of
departure
for
further
calculations.

Table
8­
1.
Incidence
of
Renal
Tubular
Regenerative
Response
in
Mice
Treated
with
HCBD
for
13
Weeks.

Test
Organism
Administered
Dose
in
Feed
(
ppm)
Approximate
Daily
Dose
(
mg/
kg­
day)
Renal
Tubule
Regeneration
Incidence
Male
Mice
0
0
0/
10
(
0%)

1
0.1
0/
10
(
0%)

3
0.4
0/
10
(
0%)

10
1.5
0/
9
(
0%)

30
4.9
10/
10
(
100%)

100
16.8
10/
10
(
100%)

Female
Mice
0
0
0/
10
(
0%)

1
0.2
1/
10
(
10%)

3
0.5
9/
10
(
90%)

10
1.8
10/
10
(
100%)

30
4.5
10/
10
(
100%)*

100
19.2
10/
10
(
100%)*

*
Not
used
in
BMD
analysis.
source:
NTP
(
1991).
5
Akaike's
Information
Criterion
(
AIC)
=
­
2L
+
2p,
where
L
is
the
log­
likelihood
at
the
maximum
likelihood
estimates
for
the
parameters,
and
p
is
the
number
of
model
degrees
of
freedom.
This
can
be
used
to
compare
models
with
different
numbers
of
parameters
using
a
similar
fitting
method
(
for
example,
least
squares
or
a
binomial
maximum
likelihood).
Although
such
methods
are
not
exact,
they
can
provide
useful
guidance
in
model
selection.

HCBD
 
February
2003
8­
3
Table
8­
2.
Benchmark
Dose
Estimates
from
NTP
(
1991)
Female
Mouse
Renal
Tubular
Regeneration
Response.

Model
BMD
BMDL
Chisquare
p­
value
AIC5
Weibull
0.200
0.099
1.00
17
Gamma
0.200
0.111
1.00
17
Log­
Probit
0.200
0.126
1.00
17
Log­
Logistic
0.200
0.123
1.00
17
Probit
0.203
0.112
0.99
17
Logistic
0.209
0.119
0.97
17
Multistage
(
2)
0.123
0.056
0.68
17
Quantal­
Quadratic
0.123
0.094
0.68
17
Quantal­
Linear
0.042
0.026
0.14
22
HCBD
 
February
2003
8­
4
Figure
8­
1.
Benchmark
Dose
Estimate
Using
Weibull
Model.
6
A
single
generation
reproductive/
developmental
study
misses
important
windows
of
vulnerability,
lactational
exposure,
and
latent
effects
that
only
become
evident
as
the
F1
generation
reaches
maturity.
That
is
the
rationale
for
the
uncertainty
factor.

HCBD
 
February
2003
8­
5
different
areas
of
uncertainty.
A
composite
uncertainty
factor
(
UF)
of
300
was
used
in
the
derivation
of
the
RfD.
The
composite
UF
included
a
factor
of
10
to
account
for
extrapolation
from
animals
to
humans;
a
factor
of
10
for
protection
of
sensitive
subpopulations;
and
a
factor
of
3
for
database
deficiencies
(
lack
of
a
2­
generation
reproductive
study6,
and
developmental
toxicity
studies
in
only
one
species).
Although
some
studies
suggest
that
humans
may
have
lower
rates
of
formation
of
putatively
toxic
metabolites
than
rodents,
others
suggest
that
the
rate
is
the
same
(
Section
6.3).

Calculation
of
RfD
Using
the
BMDL
of
0.1
mg/
kg­
day
from
the
NTP
(
1991)
study,
the
RfD
is
derived
as
follows:

RfD
=
(
0.1
mg/
kg­
day)
=
3
×
10­
4
mg/
kg­
day
300
where:

0.1
mg/
kg­
day
=
BMDL,
based
on
the
histopathological
effects
in
kidneys
of
mice
exposed
to
HCBD
in
the
diet
for
up
to
24
months
(
NTP,
1991).

300
=
uncertainty
factor.
This
is
based
on
a
factor
of
10
to
account
for
extrapolation
from
animals
to
humans;
a
factor
of
10
for
protection
of
potentially
sensitive
human
subpopulations;
and
a
factor
of
3
for
database
deficiencies
(
lack
of
a
two­
generation
reproductive
study).

8.1.2.
RfC
Determination
RfC
for
HCBD
is
not
derived.
No
subchronic
or
chronic
inhalation
exposure
studies
are
available
for
the
determination
of
RfC.

8.2
Dose­
Response
for
Cancer
Effects
8.2.1
Choice
of
Study
As
noted
previously,
only
one
lifetime
oral
carcinogenicity
study
of
HCBD
was
located
(
Kociba
et
al.,
1977).
In
this
study,
Sprague­
Dawley
rats
(
40
animals/
sex/
dose
group
and
90
animals/
sex
in
the
control
group)
were
dosed
with
0,
0.2,
2
or
20
mg/
kg­
day
HCBD
via
the
diet
for
22
months
(
males)
or
24
months
(
females).

Neoplastic
changes
were
found
only
at
the
highest
dose,
which
exceeded
the
maximum
tolerated
dose.
There
was
a
significant
increase
in
mortality
in
males,
a
greater
than
10%
decrease
in
body
weights
for
both
sexes,
and
other
severe
renal
toxicity
effects
were
observed.
The
incidence
HCBD
 
February
2003
8­
6
of
renal
tubular
neoplasms
was
increased
only
in
the
high­
dose
group
of
both
males
and
females,
as
shown
in
Table
8­
3.

Table
8­
3.
Incidence
of
Renal
Tubular
Neoplasms
in
Rats
Treated
with
HCBD
for
2
Years.

Test
Organism
Administered
Dose
(
mg/
kg­
day)
Human
Equivalent
Dosea
(
mg/
kg­
day)
Renal
Tubular
Neoplasm
Incidence
Male
rats
0
0
1/
90
(
1.1%)

0.2
0.062
0/
40
(
0%)

2.0
0.62
0/
40
(
0%)

20
5.8
9/
39
(
23%)

Female
rats
0
0
0/
90
(
0%)

0.2
0.054
0/
40
(
0%)

2.0
0.55
0/
40
(
0%)

20
5.3
6/
40
(
15%)

a
Human
Equivalent
Dose
=
Animal
dose
(
Animal
body
weight/
Human
body
weight)
1/
4
source:
Kociba
et
al.
(
1977)

Increased
renal
tubular
hyperplasia
and
renal
tubular
adenomas
and
adenocarcinomas
(
some
of
which
metastasized
to
the
lungs),
were
found
in
rats
exposed
to
20
mg/
kg­
day
of
HCBD
for
up
to
2
years.
Lesser
degrees
of
toxicity,
including
an
increase
in
renal
tubular
hyperplasia,
were
found
in
rats
ingesting
2
mg/
kg­
day
for
up
to
2
years.
A
composite
dose­
related
change
in
the
rodent
kidney
leading
to
tumor
formation
is
shown
in
Table
8­
4.
This
pattern
is
consistent
with
the
hypothesis
that
renal
tumor
formation
may
require,
and
be
secondary
to,
renal
cytotoxicity
induced
by
exposure
to
HCBD.
The
mode
of
action
established
for
HCBD
implies
that,
following
exposures
less
than
those
which
produce
overt
renal
damage,
no
significant
excess
carcinogenic
risk
can
be
attributed
to
HCBD.

8.2.2
Dose­
Response
Characterization
The
Kociba
et
al.
(
1977)
and
NTP
(
1991)
studies
were
used
to
quantify
the
cancer
risk
from
ingested
HCBD,
as
discussed
below.

1986
Guidance:
Linearized
Multistage
Model
The
current
IRIS
file
contains
a
carcinogenicity
assessment
of
HCBD
based
on
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1986).
The
dose­
response
data
for
male
rats
were
fitted
to
the
linearized
multistage
model.
To
estimate
human
equivalent
dose
from
an
animal
HCBD
 
February
2003
8­
7
Table
8­
4.
Dose­
Related
Changes
in
the
Rodent
Kidney
after
Oral
Exposure
to
HCBD,
Chronic
Study
­
Rat
(
Kociba
et
al.,
1977).

Dose
(
mg/
kg­
day)
0.2
2
20
terminal
kidney
weight
increase
(
abs.
&
rel.)
_
_
+

hyperplasia
­
multi
focal
_
?
+

hyperplasia­
adenomatous
_
+
(&
only)
+

tumors
_
_
+

study,
the
doses
administered
to
animals
were
adjusted
by
a
scaling
factor
of
(
body
weight)
2/
3.
The
resulting
cancer
slope
factor
is
7.8
×
10­
2
(
mg/
kg­
day)­
1.
This
slope
factor
corresponds
to
a
drinking
water
unit
risk
of
2.2
×
10­
6
per
:
g/
L
(
U.
S.
EPA,
1997a),
and
the
drinking
water
concentration
that
corresponds
to
a
lifetime
excess
cancer
risk
of
1
×
10­
6
is
0.5
:
g/
L.

1996,
1999
Proposed
Guidance
The
draft
Ambient
Water
Quality
Criteria
for
hexachlorobutadiene
(
U.
S.
EPA,
1998a)
utilized
the
methodology
discussed
in
EPA's
1996
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1996a)
to
evaluate
the
carcinogenicity
of
HCBD.
This
has
been
updated
in
the
1999
draft
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999).
Under
these
guidelines,
two
approaches
can
be
used
for
dose­
response
extrapolation
for
quantification
of
cancer
risk,
depending
on
what
is
known
about
the
mode
of
action
for
carcinogenicity
and
the
shape
of
the
dose­
response
curve.
A
linear
approach
is
used
for
a
chemical
when
available
evidence
indicates
the
chemical
has
direct
DNA
mutagenic
activity
or
is
DNA­
reactive,
or
when
the
evidence
supports
another
mode
of
action
that
is
anticipated
to
be
linear.
An
inference
of
linearity
may
also
be
supported
if
existing
human
exposure
is
high
and
near
doses
associated
with
key
events
in
the
carcinogenesis
process.
The
linear
approach
is
used
as
a
matter
of
policy
if
there
is
an
absence
of
sufficient
mode­
of­
action
information
on
tumorigenesis.
The
nonlinear
approach
may
be
used
when
the
tumor
mode­
of­
action
supports
nonlinearity
(
e.
g.,
some
cytotoxic
and
hormonal
agents)
and
the
chemical
does
not
demonstrate
mutagenic
effects
consistent
with
linearity.
The
nonlinear
approach
is
also
selected
when
a
mode
of
action
supporting
nonlinearity
has
been
demonstrated,
and
the
chemical
has
some
indication
of
mutagenic
activity,
but
is
judged
not
to
play
a
significant
role
in
tumor
causation.
As
a
matter
of
science
policy,
nonlinear
probability
functions
are
not
fitted
to
tumor
response
data
to
extrapolate
quantitative
low­
dose
risk
estimates
because
different
models
can
lead
7
This
modeling
was
carried
out
using
the
Global
86
multistage
model
software.

HCBD
 
February
2003
8­
8
to
a
wide
range
of
results,
and
there
is
currently
no
basis
to
choose
among
them.
In
these
cases,
the
RfD
is
generally
used
for
protection
of
cancer
effects.

Because
both
linear
(
mutagenic)
and
nonlinear
(
toxicity
associated)
mode
of
action
for
carcinogenicity
of
HCBD
may
be
operating
in
vivo,
both
of
these
approaches
were
evaluated
in
the
draft
Ambient
Water
Quality
Criteria
for
hexachlorobutadiene
(
U.
S.
EPA,
1998a)
for
characterizing
the
carcinogenic
hazard
of
HCBD,
as
discussed
below.
DNA
adduct
formation
in
mouse
kidney
was
weak,
even
at
an
oral
dose
of
30
mg/
kg.
However,
hypertrophy
and
regeneration
of
renal
tubule
epithelial
cells
precede
neoplastic
effects
in
both
dose
(
0.2
vs.
20
mg/
kg­
day)
and
time
(
subchronic
vs.
chronic
responses).
This
suggests
that
nonlinear
effects
may
underlie
or
contribute
to
the
carcinogenic
potential
of
HCBD,
although
a
role
for
genotoxic
mechanisms
cannot
confidently
be
eliminated.
Based
on
these
considerations,
the
linear
method
for
quantitative
cancer
risk
assessments
of
HCBD
would
be
the
default
approach.

Linear
Approach
Because
there
are
limited
data
which
suggest
that
HCBD
might
be
genotoxic
and
mutagenic
(
see
Section
7.3.1),
this
approach
is
considered
in
the
dose­
response
extrapolation
for
HCBD.

Under
the
proposed
guidelines,
the
cancer
risk
from
a
chemical
is
assessed
in
two
steps.
The
first
step
involves
curve­
fitting
of
the
cancer
dose­
response
data
within
the
observable
range
to
derive
a
point­
of­
departure
(
Pdp)
(
U.
S.
EPA,
1999).
The
dose
at
the
point­
of­
departure
is
expressed
as
the
human
equivalent
dose.
The
dose
that
causes
a
10%
increase
in
extra
risk
is
referred
to
as
the
ED
10.
The
point­
of­
departure
is
defined
as
the
95%
lower
confidence
limit
on
the
ED
10,
and
is
referred
to
as
the
LED
10.
The
second
step
in
the
process
is
linear
extrapolation
of
the
dose­
response
curve
from
the
LED
10
to
the
origin,
and
determination
of
the
slope
of
that
line.

The
LED
10
for
HCBD
was
calculated
by
fitting
the
quantal
polynomial
model7
to
the
tumor
dose
response
data
reported
by
Kociba
et
al.
(
1977).
Since
the
mortality
rate
was
significantly
increased
in
the
male
rats
exposed
at
the
high
dose
(
which
is
the
only
dose
with
an
increased
tumor
incidence
in
animals),
the
tumor
data
from
the
female
rats
were
used.
In
accordance
with
current
guidance
(
U.
S.
EPA,
1992d,
1999),
the
human
equivalent
dose
was
calculated
by
assuming
dose
equivalency
based
on
body
weight
raised
to
the
3/
4
power.
The
best
fit
to
the
data
is
shown
in
Figure
8­
2.
The
ED
10
was
found
to
be
4.9
mg/
kg­
day,
and
the
LED
10
was
2.5
mg/
kg­
day.
Linear
extrapolation
from
the
LED
10
to
the
origin
yields
a
slope
of
4
×
10­
2
(
mg/
kg­
day)­
1.
HCBD
 
February
2003
8­
9
0.00
0.05
0.10
0.15
0.20
0.25
0.30
0
1
2
3
4
5
6
7
Human
Equivalent
Dose
(
mg/
kg­
day)
Extra
Risk
ED10
Best
Fit
Line
Data
Points
Lower
confidence
limit
on
ED10
(
LED10)
Figure
8­
2.
Renal
Tumor
Dose
Response
Curves.

Non­
linear
Approach
The
non­
linear
approach
is
used
when
available
data
indicate
that
the
dose­
response
curve
for
tumor
induction
is
not
linear,
and
that
cancer
may
not
be
the
result
of
a
direct
DNA­
damage
mechanism.
As
discussed
previously,
data
from
the
study
by
Kociba
et
al.
(
1977)
indicate
that
the
dose
response
curve
is
strongly
non­
linear,
and
that
renal
tumors
only
occur
at
HCBD
doses
that
cause
frank
renal
toxicity
and
increased
mortality.
Therefore,
tumor
data
from
this
study
are
not
considered
suitable
for
dose­
response
extrapolation,
and
the
non­
linear
approach
should
be
evaluated.

For
HCBD,
mode­
of­
action
considerations
suggest
carcinogenicity
may
be
secondary
to
renal
toxicity,
which
has
a
threshold
(
see
Section
7.4.3,
Mode
of
Action),
and
the
RfD
approach
was
used
for
non­
linear
analysis
in
accordance
with
the
proposed
cancer
guidelines
(
U.
S.
EPA,
1999).
The
point­
of­
departure
(
Pdp)
selected
for
use
was
the
BMDL
for
renal
tubular
histological
lesions
in
female
mice.
This
is
because
a
Pdp
based
on
a
sensitive
key
precursor
of
the
neoplastic
response
is
more
protective
and
more
reliable
than
a
Pdp
based
on
the
neoplastic
response
itself
(
U.
S.
EPA,
HCBD
 
February
2003
8­
10
1999).
The
BMDL
for
renal
toxicity
calculated
from
the
NTP
(
1991)
study
was
0.1
mg/
kg­
day
in
female
mice.
Application
of
a
composite
uncertainty
factor
of
300
(
10
for
extrapolation
from
animals
to
humans;
10
for
human
variability;
and
3
for
database
deficiencies)
yields
a
dose
of
3
×
10­
4
mg/
kg­
day,
which
is
the
RfD
for
HCBD.
This
RfD
is
also
protective
of
potential
carcinogenic
effects
under
the
non­
linear
approach.

By
the
non­
linear
approach,
a
health
reference
level
(
HRL)
of
2
:
g/
L
can
also
be
derived
using
the
Reference
Dose
(
RfD)
for
HCBD
of
3
×
10­
4
mg/
kg­
day.
The
RfD
is
an
estimate
of
the
daily
oral
dose
to
the
human
population
that
is
likely
to
be
without
appreciable
risk
of
adverse
effects
over
a
lifetime
exposure.
This
dose
was
converted
to
a
drinking
water
equivalent
concentration
of
10
:
g/
L
by
multiplying
the
RfD
by
the
default
body
weight
for
an
adult
(
70
kg)
and
dividing
the
result
by
the
default
daily
intake
of
drinking
water
for
an
adult
(
2
L/
day).
For
derivation
of
the
HRL,
it
was
assumed
that
about
20%
of
an
individual's
total
exposure
to
HCBD
was
attributable
to
drinking
water.
Multiplication
of
the
drinking
water
equivalent
concentration
by
0.2
yields
the
HRL
of
2
:
g/
L
(
rounded
to
1
significant
number).
The
HRL
was
derived
as
follows:

HRL
=
RfD
×
BW
×
RSC
DI
Where:

RfD
=
Reference
dose
for
HCBD
in
drinking
water,
3
×
10­
4
mg/
kgday
BW
=
Body
weight
of
an
adult,
70
kg
DI
=
Daily
intake
of
water
for
an
adult,
2
L/
day
RSC
=
Relative
Source
Contribution,
default
value
of
20%

Therefore:

HRL
=
(
3
×
10­
4
mg/
kg­
day)
×
(
70
kg)
×
0.20
2
L/
day
=
2
:
g/
L
(
rounded
to
1
significant
number).

However,
the
HRL
of
0.9
:
g/
L
derived
from
linear
approach
is
used
as
the
preliminary
health
effect
level
in
this
document.
(
See
Section
8.2.5).
The
linear
approach
is
used
as
the
default
because
of
potential
genotoxicity
of
HCBD
metabolites.

8.2.3
Extrapolation
Model
and
Rationale
In
Section
8.2.2.,
the
carcinogenicity
of
HCBD
was
evaluated
using
both
linear
and
nonlinear
approaches.
Because
of
the
lack
of
data,
it
is
not
certain
which
method
of
cancer
risk
evaluation
is
most
appropriate
for
HCBD.
On
the
one
hand,
some
tests
indicate
that
one
or
more
of
the
metabolites
of
HCBD
are
mutagenic,
suggesting
direct
damage
to
renal
mitochondrial
DNA
by
its
reactive
metabolites.
On
the
other
hand,
direct
observations
on
cancer
dose­
response
clearly
HCBD
 
February
2003
8­
11
support
a
nonlinear
curve,
with
no
observable
increase
in
tumors
at
doses
that
do
not
induce
significant
renal
necrosis
and
regeneration.
This
is
supported
by
the
observation
that
tumors
occur
only
in
the
kidney
and
not
in
other
tissues
that
are
not
significantly
injured
by
HCBD.
Therefore,
the
tumor
data
from
the
Kociba
et
al.
(
1977)
study
are
not
considered
suitable
for
linear
dose­
response
extrapolation.

Although
HCBD
metabolites
have
some
indication
of
mutagenic
activity,
they
are
not
likely
to
play
a
significant
role
in
tumor
causation
due
to
their
weak
activity.
Moreover,
mutations
of
mitochondrial
DNA
may
result
in
mitochondrial
dysfunction
(
See
Section
7.4.3),
which
would
support
cytotoxicity
and
nonlinear
approach.
There
should
also
be
decreased
concern
over
genotoxicity
for
humans
because
the
activity
of
HCBD
metabolizing
enzymes,
particularly
renal
$­
lyase,
may
be
many
fold
lower
in
humans
than
the
corresponding
enzymes
in
rats
(
see
section
6.3).
In
addition,
human
exposure
levels
(
see
Section
9.3)
are
about
4
orders
of
magnitude
lower
than
the
human
equivalent
dose
corresponding
to
the
dose
at
which
tumor
incidence
was
reported
in
Kociba
et
al.
(
1977)
study.
In
consideration
of
the
overall
evidence,
the
non­
linear
approach
may
be
appropriate
for
HCBD.
The
draft
Ambient
Water
Quality
Criteria
Document
for
Hexachlorobutadiene
(
U.
S.
EPA,
1998a)
has
recommended
using
the
nonlinear
approach
for
carcinogenicity
assessment
of
HCBD.

According
to
the
1999
U.
S.
EPA
Draft
Guidelines
for
Carcinogen
Risk
Assessment,
"
When
the
mode
of
action
information
indicates
that
the
dose­
response
may
be
adequately
described
by
both
a
linear
and
a
nonlinear
approach,
then
the
default
is
to
present
both
the
linear
and
margin
of
exposure
analyses."
As
this
can
be
done
without
contradicting
previous
guidelines,
both
linear
and
nonlinear
analyses
are
presented.

8.2.4
Cancer
Potency
and
Unit
Risk
Table
8­
5
summarizes
the
cancer
values
derived
for
HCBD.
Analysis
of
tumor
dose­
response
information
from
the
Kociba
et
al.
(
1977)
study
using
the
linear
extrapolation
approach
from
the
proposed
carcinogen
risk
assessment
guidelines
(
U.
S.
EPA,
1999)
resulted
in
a
slope
factor
of
4
×
10­
2
(
mg/
kg­
day)­
1.
This
value
is
about
half
of
the
slope
factor
of
7.8
×
10­
2
(
mg/
kg­
day)­
1
derived
previously
using
the
linearized
multistage
(
LMS)
model
(
U.
S.
EPA,
1997a),
but
most
of
the
apparent
difference
may
be
attributable
to
the
different
methods
used
to
calculate
human
equivalent
doses
from
the
animal
doses
(
the
scaling
factor
used
in
the
LMS
approach
assumed
body
weight
to
the
2/
3
power,
while
a
factor
of
body
weight
raised
to
the
3/
4
was
used
for
the
Pdp
method).
Based
on
the
slope
factor
of
4
×
10­
2
(
mg/
kg­
day)­
1
derived
using
the
LED
10
approach
with
linear
extrapolation,
the
unit
risk
is
1.1
×
10­
6
per
(:
g/
L)
and
the
drinking
water
concentration
that
corresponds
to
a
lifetime
excess
risk
of
1
×
10­
6
is
0.9
:
g/
L.
HCBD
 
February
2003
8­
12
Table
8­
5.
Summary
of
Cancer
Risk
Values
for
HCBD.

Approach
Parameter
Value
LMSa
(
U.
S.
EPA,
1991)
Slope
7.8
×
10­
2
(
mg/
kg­
day)­
1
Unit
Risk
2.2
×
10­
6
per
(:
g/
L)

Water
Concentration
at
risk
of
1
×
10­
6
0.5
:
g/
L
LED
10,
linear
extrapolation
LED
10
(
tumors)
b
2.5
mg/
kg­
day
Slope
4
×
10­
2
(
mg/
kg­
day)­
1
Unit
Risk
1.1
×
10­
6
per
(:
g/
L)

Water
Concentration
at
risk
of
1
×
10­
6
(
HRL)
0.9
:
g/
L
Nonlinear
Pdp
(
BMDL)
0.1
mg/
kg­
day
Uncertainty
factor
300
HRL
2
:
g/
L
a
Animal
to
human
dose
extrapolation
based
on
body
weight2/
3
b
Animal
to
human
dose
extrapolation
based
on
body
weight3/
4
8.2.5
Discussion
of
Confidence
The
available
database
associating
HCBD
and
carcinogenicity
is
limited.
There
are
no
human
data.
The
evidence
is
obtained
only
in
one
chronic
dietary
study
in
a
single
species
(
Sprague­
Dawley
rats)
(
Kociba
et
al.,
1977),
where
rats
developed
severe
renal
toxicity
preceding
tumor
formation.
The
tumors
were
seen
only
at
a
high
dose
which
exceeded
the
maximum
tolerated
dose
(
MTD,
i.
e.,
greater
than
10%
body
weight
depression)
in
both
sexes
of
rats
and
produced
high
mortality
in
the
males.
Similar
renal
toxicity
observed
in
a
30­
day
study
of
HCBD
in
rats
by
the
same
laboratory
and
in
another
90­
day
subchronic
study
in
mice
(
NTP,
1991)
strengthens
the
idea
that
the
tumor
formation
is
induced
by
cytotoxicity.
Both
the
NTP
(
1991)
and
Kociba
et
al.
(
1977)
studies
tested
a
sufficient
number
of
animals.

A
limitation
of
the
Kociba
et
al.
(
1977)
study
is
the
selection
and
spacing
of
doses.
Although
the
study
employed
an
adequate
number
of
animals,
the
doses
selected
for
testing
were
separated
by
a
factor
of
10
(
0,
0.2,
2,
and
20
mg/
kg­
day).
Thus,
there
are
no
observations
between
the
dose
of
2
mg/
kg­
day
(
causing
no
tumors),
and
the
dose
of
20
mg/
kg­
day
(
causing
a
15%
tumor
response
in
females
and
a
23%
tumor
response
in
males).
More
doses
between
2
and
20
mg/
kg­
day
would
better
delineate
the
shape
of
the
dose­
response
curve.
HCBD
 
February
2003
8­
13
A
weight­
of­
evidence
analysis
of
the
available
data
as
a
whole
indicates
that
the
confidence
in
using
either
the
linear
or
nonlinear
approach
is
not
high;
this
is
particularly
true
for
the
linear
method
which
is
based
on
only
one
data
point
at
a
high­
dose
exceeding
the
MTD.
However,
the
possible
genotoxicity
of
HCBD
metabolites
forces
the
use
of
the
linear
approach
as
default.
Thus,
the
HRL
of
0.9
µ
g/
L
derived
from
the
linear
approach
is
used
as
the
preliminary
health
effect
level
in
this
document.
HCBD
 
February
2003
9­
1
9.0
REGULATORY
DETERMINATION
AND
CHARACTERIZATION
OF
RISK
FROM
DRINKING
WATER
9.1
Regulatory
Determination
for
Chemicals
on
the
CCL
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1996,
required
the
Environmental
Protection
Agency
(
EPA)
to
establish
a
list
of
contaminants
to
aid
the
Agency
in
regulatory
priority
setting
for
the
drinking
water
program.
EPA
published
a
draft
of
the
first
Contaminant
Candidate
List
(
CCL)
on
October
6,
1997
(
62
FR
52193,
U.
S.
EPA,
1997b).
After
review
of
and
response
to
comments,
the
final
CCL
was
published
on
March
2,
1998
(
63FR
10273,
U.
S.
EPA
1998d).
The
CCL
grouped
contaminants
into
three
major
categories
as
follows:

Regulatory
Determination
Priorities
­
Chemicals
or
microbes
with
adequate
data
to
support
a
regulatory
determination,

Research
Priorities
­
Chemicals
or
microbes
requiring
research
for
health
effects,
analytical
methods,
and/
or
treatment
technologies,

Occurrence
Priorities
­
Chemicals
or
microbes
requiring
additional
data
on
occurrence
in
drinking
water.

The
March
2,
1998
CCL
included
one
microbe
and
19
chemicals
in
the
regulatory
determination
priority
category.
More
detailed
assessments
of
the
completeness
of
the
health,
treatment,
occurrence
and
analytical
method
data
led
to
a
subsequent
reduction
of
the
regulatory
determination
priority
chemicals
to
a
list
of
12
(
one
microbe
and
11
chemicals)
which
was
distributed
to
stakeholders
in
November
1999.

SDWA
requires
EPA
to
make
regulatory
determinations
for
no
fewer
than
five
contaminants
in
the
regulatory
determination
priority
category
by
August
2001.
In
cases
where
the
Agency
determines
that
a
regulation
is
necessary,
the
regulation
should
be
proposed
by
August
2003
and
promulgated
by
February
2005.
The
Agency
is
given
the
freedom
to
also
determine
that
there
is
no
need
for
a
regulation
if
a
chemical
on
the
CCL
fails
to
meet
one
of
three
statutory
criteria
established
by
SDWA
and
described
in
Section
9.1.1.

9.1.1
Criteria
for
Regulatory
Determination
These
are
the
three
criteria
used
to
determine
whether
or
not
to
regulate
a
chemical
on
the
CCL:

The
contaminant
may
have
an
adverse
effect
on
the
health
of
persons,

The
contaminant
is
known
to
occur,
or
there
is
a
substantial
likelihood
that
the
contaminant
will
occur,
in
public
water
systems
with
a
frequency
and
at
levels
of
public
health
concern,
HCBD
 
February
2003
9­
2
In
the
sole
judgment
of
the
administrator,
regulation
of
such
contaminant
presents
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.

The
findings
for
all
criteria
are
used
in
making
a
determination
to
regulate
a
contaminant.
As
required
by
SDWA,
a
decision
to
regulate
commits
the
EPA
to
publication
of
a
Maximum
Contaminant
Level
Goal
(
MCLG)
and
promulgation
of
a
National
Primary
Drinking
Water
Regulation
(
NPDWR)
for
that
contaminant.
The
agency
may
determine
that
there
is
no
need
for
a
regulation
when
a
contaminant
fails
to
meet
one
of
the
criteria.
A
decision
not
to
regulate
a
contaminant
is
considered
a
final
Agency
action
and
is
subject
to
judicial
review.
The
Agency
can
choose
to
publish
a
Health
Advisory
(
a
nonregulatory
action)
or
other
guidance
for
any
contaminant
on
the
CCL
independent
of
the
regulatory
determination.

9.1.2
National
Drinking
Water
Advisory
Council
Recommendations
In
March
2000,
the
EPA
convened
a
Working
Group
under
the
National
Drinking
Water
Advisory
Council
(
NDWAC)
to
help
develop
an
approach
for
making
regulatory
determinations.
The
Working
Group
developed
a
protocol
for
analyzing
and
presenting
the
available
scientific
data,
and
recommended
methods
to
identify
and
document
the
rationale
supporting
a
regulatory
determination
decision.
The
NDWAC
Working
Group
report
was
presented
to
and
accepted
by
the
entire
NDWAC
in
July
2000.

Because
of
the
intrinsic
difference
between
microbial
and
chemical
contaminants,
the
Working
Group
developed
separate
but
similar
protocols
for
microorganisms
and
chemicals.
The
approach
for
chemicals
was
based
on
an
assessment
of
the
impact
of
acute,
chronic,
and
lifetime
exposures,
as
well
as
a
risk
assessment
that
includes
evaluation
of
occurrence,
fate,
and
doseresponse
The
NDWAC
protocol
for
chemicals
is
a
semi­
quantitative
tool
for
addressing
each
of
the
three
CCL
criteria.
The
NDWAC
requested
that
the
Agency
use
good
judgement
in
balancing
the
many
factors
that
need
to
be
considered
in
making
a
regulatory
determination.

The
EPA
modified
the
semi­
quantitative
NDWAC
suggestions
for
evaluating
chemicals
against
the
regulatory
determination
criteria
and
applied
them
in
decision
making.
The
quantitative
and
qualitative
factors
for
hexachlorobutadiene
(
HCBD)
that
were
considered
for
each
of
the
three
criteria
are
presented
in
the
sections
that
follow.

9.2
Health
Effects
The
first
criterion
asks
if
the
contaminant
may
have
an
adverse
effect
on
the
health
of
persons.
Because
all
chemicals
have
adverse
effects
at
some
level
of
exposure,
the
challenge
is
to
define
the
dose
at
which
adverse
health
effects
are
likely
to
occur,
and
estimate
a
dose
at
which
adverse
health
effects
are
either
not
likely
to
occur
(
threshold
toxicant),
or
have
a
low
probability
for
occurrence
(
non­
threshold
toxicant).
The
key
elements
that
must
be
considered
in
evaluating
the
first
criterion
are
the
mode
of
action,
the
critical
effect(
s),
the
dose­
response
for
critical
effect(
s),
the
RfD
for
threshold
effects,
and
the
slope
factor
for
non­
threshold
effects.
HCBD
 
February
2003
9­
3
A
description
of
the
health
effects
associated
with
exposure
to
HCBD
is
presented
in
Chapter
7
of
this
document
and
summarized
below
in
Section
9.2.2.
Chapter
8
and
Section
9.2.3
present
dose­
response
information,
where
applicable,
for
threshold
and
non­
threshold
health
effects.

9.2.1
Health
Criterion
Conclusion
The
available
toxicological
data
indicate
that
HCBD
has
the
potential
to
cause
adverse
health
effects
in
animals,
and
probably
in
humans.
The
available
human
data
involve
inhalation
exposure
and
are
confounded
by
simultaneous
exposures
to
other
chemicals
in
an
occupational
setting;
thus,
attributing
observed
effects
to
specific
levels
of
HCBD
exposure
is
not
possible.
In
rodents,
there
is
clear
evidence
of
renal
damage
resulting
from
acute,
subchronic,
and
chronic
HCBD
oral
exposures.
A
few
animal
studies
have
also
reported
liver
effects
and
neurotoxicity.
Review
of
animal
dose­
response
data
endpoints
indicates
that
subchronic
and
chronic
LOAEL
values
for
HCBD
toxicity
are
generally
at
2
mg/
kg­
day
and
above.
The
RfD
for
HCBD
is
3
×
10­
4
mg/
kg­
day
(
Chapter
8).
Limited
evidence
of
carcinogenic
potential
in
rodents
suggests
that
HCBD
may
be
carcinogenic
secondary
to
renal
tubular
epithelial
cell
cytotoxicity.
However,
data
in
humans
are
lacking.
Both
linear
and
non­
linear
approaches
were
evaluated
for
cancer
dose­
response
assessment.
Using
the
non­
linear
approach,
the
RfD
protects
both
non­
cancer
and
cancer
effects.
However,
in
the
presence
of
data
supporting
the
potential
genotoxicity
of
HCBD
metabolites,
the
linear
approach
is
used
as
default,
with
a
10­
6
risk
at
a
drinking
water
concentration
of
0.9
µ
g/
L.

9.2.2
Hazard
Characterization
and
Mode
of
Action
Implications
Data
for
the
human
health
effects
of
HCBD
are
limited
to
a
few
studies
of
occupational
exposure
to
HCBD.
A
relationship
could
not
be
established
from
these
studies
between
HCBD
exposure
and
toxic
effects
either
because
of
concurrent
exposure
to
other
chemicals
or
because
of
equivocal
results.

Studies
in
animals
show
the
selective
effect
of
HCBD
on
the
kidney,
specifically
the
proximal
tubule.
Renal
toxicity
in
rodents
has
been
shown
with
single
acute
exposures
to
100
 
200
mg
HCBD/
kg,
and
with
short­
term
exposures
to
3
mg/
kg­
day
and
above.
Subchronic
and
chronic
studies
in
rodents
show
clear
dose­
related
renal
damage
at
2
mg/
kg­
day
and
above.
Progressive
events
over
time
include
changes
in
kidney
weight,
increased
urinary
excretion
of
coproporphyrin,
and
increased
renal
tubular
epithelial
hyperplasia.

Other
noncancer
effects
associated
with
HCBD
exposure
in
animals
include
developmental
effects
and
neurotoxicity
(
Harleman
and
Seinen,
1979;
Badaeva,
1983;
Badaeva
et
al.,
1985).
However,
these
effects
were
observed
at
higher
doses
than
for
renal
toxicity.
Pups
with
lower
birth
weights
and
reduced
growth
were
reported
at
maternal
dose
of
8.1­
15
mg/
kg­
day
in
rats
(
Badaeva,
1983;
Harleman
and
Seinen,
1979).

Results
from
mutagenicity
studies
with
HCBD
are
ambiguous.
In
the
presence
of
appropriate
metabolic
activation
conditions,
HCBD
and
its
metabolites
are
mutagenic
in
some
(
Vamvakas
et
al.,
1988;
Reichert
et
al.,
1984),
but
not
all
studies.
HCBD
metabolites
have
been
shown
to
bind
to
mitochondrial
DNA
in
vivo
in
mice
(
Schrenk
and
Dekant,
1989),
and
induce
DNA
repair
in
cultured
HCBD
 
February
2003
9­
4
porcine
kidney
cells
(
Vamvakas
et
al.,
1989),
suggesting
its
genotoxic
potential.
No
human
studies
of
HCBD
carcinogenicity
have
been
reported
and
only
one
lifetime
animal
study
has
been
performed
(
Kociba
et
al.,
1977).
In
this
study,
neoplastic
changes
occurred
only
at
the
highest
dose
which
exceeded
the
maximum
tolerated
dose
(
MTD),
i.
e.
there
was
increased
mortality,
greater
than
10%
decrease
in
body
weight
and
severe
renal
toxicity.
Because
these
significant
adverse
effects
were
observed
at
the
high
dose,
tumor
formation
may
be
secondary
to
cytotoxicity.

The
nephrotoxicity
of
HCBD
is
dependent
on
a
multistep
bioactivation
mechanism
involving
both
kidney
and
liver
enzymes.
The
ultimate
step
in
this
pathway
is
a
$­
lyase
mediated
degradation
of
a
HCBD
metabolite
that
generates
a
highly
reactive
thioketene
in
proximal
tubule
cells..
In
vitro
studies
suggest
that
cortical
mitochondria
are
the
critical
subcellular
target
for
toxicity
of
the
bioactivated
sulfur
conjugates
of
HCBD.
Covalent
binding
of
this
reactive
HCBD
metabolite
to
cellular
macromolecules
(
e.
g.
proteins,
mitochondrial
DNA),
and
the
resultant
mitochondrial
dysfunction
is
believed
to
contribute
to
the
renal
cytotoxicity
and
tumors
observed
in
animals.
Recent
evidence
suggests
that
HCBD­
induced
"
2:­
globulin
accumulation
contributes
to
renal
injury
in
male
rats.
However,
renal
tubular
necrosis
and
renal
tubular
tumors
were
observed
in
both
male
and
female
rats
following
HCBD
exposure
(
Kociba
et
al.,
1977),
and
renal
necrosis
and
regeneration
were
also
observed
in
male
and
female
mice
(
NTP,
1991).
Therefore,
"
2:­
globulin
accumulation
cannot
be
the
sole
mechanism
for
HCBD­
induced
carcinogenesis.

One
important
issue
in
the
evaluation
of
the
hazard
posed
by
HCBD
is
the
applicability
of
rodent
mechanistic
data
to
humans.
In
vitro
studies
with
human
renal
cytosol
and
cultured
human
proximal
tubule
cells
suggest
that
humans
have
the
potential
to
form
the
HCBD­
glutathione
conjugates
and
to
metabolize
HCBD
cysteine
conjugates
to
toxic
metabolites.
However,
the
rate
of
metabolism,
particularly
for
the
reaction
catalyzed
by
$­
lyase,
appears
to
be
much
lower
for
humans
than
rodents
(
Lock,
1994;
Lash
et
al.,
1990).

It
has
been
generally
observed
that
existing
nephropathy
or
age­
related
kidney
degeneration
can
increase
the
risk
of
renal
injury
or
exacerbate
nephrotoxicity
in
humans.
Therefore,
sensitive
populations
for
HCBD
exposure
may
include
people
with
pre­
existing
kidney
or
liver
damage
or
the
elderly.
Although
it
is
unlikely
that
human
newborns
would
be
acutely
exposed
to
significant
doses
of
HCBD,
acute
exposures
for
young
rats
and
mice
cause
toxicity
at
lower
doses
than
for
adults
(
Hook
et
al.,
1983;
Lock
et
al.,
1984).

9.2.3
Dose­
Response
Characterization
and
Implications
in
Risk
Assessment
Dose­
response
information
from
several
key
studies
of
HCBD
toxicity
in
animals
is
summarized
in
Table
9­
1.
These
studies
currently
provide
the
most
reliable
information
on
threshold
levels
for
HCBD
toxicity
in
animals
exposed
via
the
oral
route.

Noncancer
effects
In
short­
term
studies,
a
LOAEL
of
10
mg/
kg­
day
and
a
NOAEL
of
3
mg/
kg­
day
were
identified
for
reduced
body
weight
gain
and
food
consumption
in
female
Sprague­
Dawley
rats
administered
HCBD
in
their
diets
for
30
day.
Renal
tubular
degeneration,
necrosis
and
regeneration
HCBD
 
February
2003
9­
5
Table
9­
1.
Dose­
Response
Information
from
Several
Key
Studies
of
HCBD
Toxicity
(
Oral
Exposure).

Study
Species
No./
Sex
Doses
mg/
kg­
day
Duration
NOAEL
mg/
kgday
LOAEL
mg/
kgday
Effects
Short­
term
Studies
Kociba
et
al.
(
1977)
Rat
Sprague­
Dawley
4
F
1.310e+
10
30
days
3
10
Reduced
body
weight
gain,
food
consumption;
increased
hemoglobin
concentration,
relative
kidney
weight;
renal
tubular
degeneration,
necrosis,
regeneration.

Jonker
et
al.
(
1993b)
Rat
Wistar
5
M
5
F
2.25828
4
weeks
2.25
8
Decreased
liver
weight,
plasma
creatinine,
body
weight,
adrenal
weight;
renal
tubular
cytomegaly.

Harleman
and
Seinen
(
1979)
Rat
Wistar
6
M
6
F
0
4.6
14.0
35.3
14
days
­­
4.6
Decreased
body
weight
gain
and
food
conversion
efficiency;
renal
tubular
epithelial
cell
degeneration.

Stott
et
al.
(
1981)
Rat
Sprague­
Dawley
5
M
0.22
3
weeks
0.2
20
Decreased
body
weight
gain;
increased
relative
kidney
weight;
kidney
damage.

NTP
(
1991)
Mouse
B6C3F
1
5
M
5
F
0
M
0
F
3
M
5
F
12
M
16
F
40
M
49
F
2
weeks
­­
3
M
5
F
Renal
tubular
necrosis.
Table
9­
1
(
continued)

Study
Species
No./
Sex
Doses
mg/
kg­
day
Duration
NOAEL
mg/
kgday
LOAEL
mg/
kgday
Effects
HCBD
 
February
2003
9­
6
Subchronic
Studies
NTP
(
1991)
Mouse
B6C3F
1
10
M
10
F
0
M
0
F
0.1
M
0.2
F
0.4
M
0.5
F
1.5
M
1.8
F
4.9
M
4.5
F
16.8
M
19.2
F
13
weeks
1.5
M
0.2
F
4.9
M
0.5
F
Renal
tubular
cell
regeneration
(
increased
epithelial
nuclei
and
basophilic
staining)

Chronic
Studies
Kociba
et
al.
(
1977)
Rat
Sprague­
Dawley
39
 
40
M,
F
0.222
22
 
24
months
0.2
2
Increased
kidney
weight;
renal
tubular
epithelial
hyperplasia
and
neoplasia.

M
=
male;
F
=
female
were
observed
at
30
mg/
kg­
day
(
Kociba
et
al.,
1971;
Schwetz
et
al.,
1977).
A
LOAEL
of
8
mg/
kgday
and
a
NOAEL
of
2.25
mg/
kg­
day
were
identified
for
decreased
body
weight
gain
and
renal
tubular
effects
in
Wistar
rats
given
HCBD
in
their
diets
for
4
weeks
(
Jonker
et
al.,
1993b).
A
3­
week
oral
exposure
with
male
Sprague­
Dawley
rats
identified
a
LOAEL
of
20
mg/
kg­
day
and
a
NOAEL
of
0.2
mg/
kg­
day
for
kidney
damage
and
increased
relative
kidney
weight
(
Stott
et
al.,
1981),
and
a
2­
week
feeding
study
in
Wistar
rats
identified
a
LOAEL
of
4.6
mg/
kg­
day
(
the
lowest
dose
tested)
for
renal
tubular
epithelial
cell
degeneration
(
Harleman
and
Seinen,
1979).
A
2­
week
oral
exposure
study
in
B6C3F
1
mice
reported
a
LOAEL
of
3
 
5
mg/
kg­
day
(
the
lowest
dose
tested)
for
renal
tubular
necrosis
(
NTP,
1991).
Thus,
renal
effects
in
rodents
resulting
from
short­
term
exposure
to
HCBD
appear
to
have
LOAELs
of
around
5
 
20
mg/
kg­
day,
depending
on
the
species
and
strain
used,
the
length
of
exposure,
and
the
method
of
administration.

In
a
subchronic
oral
exposure
study
of
HCBD
in
B6C3F
1
mice,
a
NOAEL
of
1.5
mg/
kg­
day
was
identified
for
male
mice
based
on
renal
tubular
cell
regeneration
(
NTP,
1991).
Tubular
regeneration
occurred
in
1
of
10
females
in
the
lowest
dose
group
(
0.2
mg/
kg­
day).
The
study
authors
concluded
that
a
NOAEL
for
female
mice
could
not
be
identified
from
these
data
(
NTP,
1991).
However,
EPA
(
U.
S.
EPA,
1998a)
and
others
(
WHO,
1994)
have
concluded
that
the
effect
observed
at
0.2
mg/
kg­
day
is
not
statistically
significant,
and
therefore
consider
this
dose
to
be
the
NOAEL
for
female
mice.
Because
tubular
regeneration
occurred
in
1
of
10
females
at
0.2
mg/
kg­
day,
this
NOAEL
may
be
close
to
a
minimal
LOAEL
for
renal
injury.

Only
one
study
of
lifetime
oral
exposure
to
HCBD
was
located
(
Kociba
et
al.,
1977).
This
study
identified
a
NOAEL
of
0.2
mg/
kg­
day
and
a
LOAEL
of
2
mg/
kg­
day
in
rats,
based
on
an
HCBD
 
February
2003
9­
7
increase
in
renal
tubular
epithelial
cell
hyperplasia/
regeneration.
The
value
of
this
NOAEL
from
a
chronic
study
is
the
same
as
the
equivocal
NOAEL
of
0.2
mg/
kg­
day
identified
in
the
13­
week
NTP
study
in
female
mice
(
NTP,
1991),
indicating
the
female
mice
may
be
more
sensitive
than
rats
to
HCBD.

The
Reference
Dose
(
RfD)
for
HCBD
is
3
×
10­
4
mg/
kg­
day
(
Chapter
8).
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
BMDL
of
0.1
mg/
kg­
day
for
renal
tubular
epithelial
cell
hyperplasia/
regeneration
from
the
NTP
(
1991)
studies.
A
composite
uncertainty
factor
of
300
was
used
in
the
derivation
of
the
RfD
to
account
for:
extrapolation
from
animals
to
humans
(
factor
of
10);
protection
of
sensitive
subpopulations
(
factor
of
10);
and
database
deficiency
(
factor
of
3)
because
of
lack
of
a
2­
generation
reproductive
study.

Cancer
effects
The
single
lifetime
exposure
study
in
rats
is
also
a
source
of
data
on
tumor
formation
(
Kociba
et
al.,
1977).
Only
at
the
highest
dose,
20
mg/
kg­
day,
were
tumors
seen
in
both
sexes.
This
dose
exceeded
the
level
at
which
significant
noncancer
effects
were
seen,
such
as
mortality,
renal
toxicity,
and
body
weight
depression.
In
this
study,
the
second
highest
dose
was
2
mg/
kg­
day
and
there
were
no
tumors
in
this
exposed
group.
While
slope
of
the
dose­
response
curve
cannot
be
determined
from
the
data
set,
it
must
be
kept
in
mind
that
the
purpose
of
the
Kociba
et
al.(
1977)
bioassay
was
for
hazard
identification,
not
quantitative
risk
analysis.

Under
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1986),
HCBD
is
classified
as
Group
C,
possible
human
carcinogen.
Using
the
linearized
multistage
model,
a
slope
factor
of
7.8
×
10­
2
per
mg/
kg­
day
was
calculated
at
the
95th
upper
confidence
level
(
U.
S.
EPA,
1991c).
Under
EPA's
1999
draft
Guidelines
for
Carcinogen
Risk
Assessment
(
USEPA,
1999),
HCBD
is
classified
as
likely
to
be
carcinogenic
to
humans
by
the
oral
route
of
exposure,
but
whose
carcinogenic
potential
by
the
inhalation
and
dermal
routes
of
exposure
cannot
be
determined
because
there
are
inadequate
data
to
perform
an
assessment.
Both
the
linear
and
nonlinear
doseresponse
extrapolation
approaches
were
used
to
quantify
cancer
risk
(
U.
S.
EPA,
1998a)
because
both
cytotoxicity
and
mutagenic
mode
of
action
may
be
involved.
The
linear
approach
yields
a
slope
of
4
×
10­
2
per
mg/
kg­
day.
Using
the
non­
linear
approach,
the
RfD
was
used,
yielding
the
resulting
dose
of
3
×
10­
4
mg/
kg­
day.
EPA's
draft
Ambient
Water
Quality
Criteria
for
hexachlorobutadiene
(
U.
S.
EPA,
1998a)
recommended
using
the
non­
linear
approach
for
dose­
response
extrapolation.
As
discussed
previously,
data
from
Kociba
et
al.
(
1977)
indicated
that
the
tumor
dose
response
curve
is
strongly
non­
linear,
and
that
renal
tumors
only
occur
at
HCBD
doses
that
cause
frank
toxicity.
However,
the
presence
of
data
supporting
the
genotoxicity
of
HCBD
metabolites
forces
the
conservative
choice
of
a
linear
approach
to
be
more
appropriate
for
HCBD.

The
conclusion
from
the
dose
response
analysis
is
that
HCBD
is
a
weak
carcinogen
because
it
is
carcinogenic
only
at
cytotoxic
dose.
HCBD
 
February
2003
9­
8
9.3
Occurrence
in
Public
Water
Systems
The
second
criterion
asks
if
the
contaminant
is
known
to
occur
or
if
there
is
a
substantial
likelihood
that
the
contaminant
will
occur
in
public
water
systems
with
a
frequency
and
at
levels
of
public
health
concern.
In
order
to
address
this
question,
the
following
information
was
considered:

°
Monitoring
data
from
public
water
systems
°
Ambient
water
concentrations
and
releases
to
the
environment
°
Environmental
fate
Data
on
the
occurrence
of
HCBD
in
public
drinking
water
systems
were
the
most
important
determinants
in
evaluating
the
second
criterion.
EPA
looked
at
the
total
number
of
systems
that
reported
detections
of
HCBD,
as
well
as
those
that
reported
concentrations
of
HCBD
above
an
estimated
drinking
water
health
reference
level
(
HRL)
(
U.
S.
EPA,
2001c).
For
noncarcinogens,
the
estimated
HRL
level
was
calculated
from
the
RfD
assuming
that
20%
of
the
total
exposure
would
come
from
drinking
water.
For
carcinogens,
the
HRL
was
the
10­
6
risk
level.
The
HRLs
are
benchmark
values
that
were
used
in
evaluating
the
occurrence
data
while
the
risk
assessments
for
the
contaminants
were
being
developed.

The
available
monitoring
data,
including
indications
of
whether
or
not
the
contamination
is
a
national
or
a
regional
problem,
are
included
in
Chapter
4
of
this
document
and
summarized
below.
Additional
information
on
production,
use,
and
fate
are
found
in
Chapters
2
and
3.

9.3.1
Occurrence
Criterion
Conclusion
HCBD
has
never
been
specifically
manufactured
as
a
commercial
product
in
the
United
States,
but
is
generated
as
waste
by­
product
from
the
chlorination
of
hydrocarbons.
The
available
data
for
HCBD
use
indicate
an
overall
downward
trend.
The
ten­
year
pattern
of
TRI
releases
to
surface
water
is
variable
but
generally
decreasing
within
the
range
from
5
to
1,911
pounds.
The
physicochemical
properties
of
HCBD
and
the
available
data
for
environmental
fate
indicate
that
HCBD
in
surface
water
is
likely
to
be
rapidly
degraded
by
biotic
and
abiotic
processes
although
it
has
the
potential
for
bioaccumulation.
Monitoring
data
indicate
that
HCBD
is
infrequently
detected
in
public
water
supplies.
When
HCBD
is
detected,
it
very
rarely
exceeds
the
HRL
or
a
value
of
onehalf
of
the
HRL.
Chemical
treatment
of
drinking
water
and
leaching
from
drinking
water
surfaces
are
not
expected
to
contribute
to
significantly
elevated
levels
of
HCBD
in
drinking
water.

9.3.2
Monitoring
Data
Drinking
Water
HCBD
has
been
detected
in
a
small
percentage
of
public
water
supply
(
PWS)
samples
collected
under
the
authority
of
the
Safe
Drinking
Water
Act.
Occurrence
data
for
HCBD
in
drinking
water
are
presented
and
analyzed
in
Chapter
4
of
this
document.
Data
from
two
monitoring
periods
HCBD
 
February
2003
9­
9
were
available
for
analysis.
Data
from
Round
1
were
collected
during
the
period
1987
to
1992.
Data
from
Round
2
were
collected
during
the
period
1993
to
1997.
Round
1
and
2
monitoring
detected
HCBD
in
only
0.13%
and
0.05%
of
all
samples
analyzed,
respectively.
When
data
are
expressed
on
a
PWS
basis,
Round
1
and
Round
2
monitoring
detected
HCBD
at
least
once
in
0.35%
(
228
systems)
and
0.18%
(
117
systems)
of
the
tested
water
supplies,
respectively.

The
median
and
99th
percentile
concentrations
for
all
samples
(
i.
e.,
samples
with
and
without
detectable
levels
of
HCBD)
were
below
the
minimum
reporting
level
(
MRL).
When
subsets
of
the
data
containing
only
samples
with
detectable
levels
of
HCBD
were
analyzed,
the
median
and
99th
percentile
concentrations
for
Round
1
were
0.25
:
g/
L
and
10
:
g/
L,
respectively.
The
median
and
99th
percentile
for
Round
2
detections
were
0.30
:
g/
L
and
1.5
:
g/
L,
respectively.

When
monitoring
results
were
compared
to
a
value
of
one­
half
of
the
HRL,
0.163%
of
Round
1
(
106
systems)
and
0.079%
of
Round
2
(
51
systems)
water
supplies
exceeded
this
benchmark
at
least
once
during
the
reporting
period.
The
percentages
of
water
supplies
that
exceeded
the
HRL
at
least
once
in
Round
1
and
Round
2
monitoring
were
0.114%
(
74
systems)
and
0.018%
(
11
systems),
respectively.

PWSs
with
detected
levels
of
HCBD
were
widely
distributed
throughout
the
United
States
(
see
Figures
4­
2
and
4­
3
in
this
document),
and
no
clear
patterns
of
regional
geographic
occurrence
were
evident.

Ambient
Water
HCBD
has
not
been
detected
in
the
ground
water
samples
reviewed
and/
or
analyzed
under
the
U.
S.
Geological
Survey
National
Ambient
Water
Quality
Assessment
(
NAWQA)
program.
The
first
round
of
intensive
monitoring
in
the
ongoing
NAWQA
was
conducted
from
1991
to
1996
and
targeted
20
watersheds.
Data
from
each
NAWQA
study
unit
were
augmented
by
additional
data
from
local,
state,
and
federal
agencies
that
met
specified
criteria.(
See
Section
4.1.1).
HCBD
was
not
detected
in
rural
and
urban
wells
of
the
local,
State,
and
federal
data
set
compiled
by
NAWQA.
These
data
represent
untreated
ground
water
of
the
conterminous
United
States
for
the
years
1985­
1995.

A
review
of
highway
and
urban
runoff
studies
also
found
no
detections
of
HCBD.

9.3.3
Use
and
Fate
Data
Significant
quantities
of
HCBD
are
generated
in
the
United
States
as
waste
by­
product
from
the
chlorination
of
hydrocarbons,
although
HCBD
has
never
been
specifically
manufactured
as
a
commercial
product
domestically.
No
recent
estimate
could
be
found
on
the
by­
product
amounts,
but
in
1982,
it
was
estimated
that
about
28
million
pounds
were
generated
(
ATSDR,
1994).
HCBD
imports
dropped
during
the
late
1970s,
the
period
for
which
data
are
reported
(
Howard,
1989).
HCBD
 
February
2003
9­
10
In
all
environmental
media,
HCBD
binds
strongly
to
particles
(
ATSDR,
1994).
It
is
readily
adsorbed
to
airborne
particulate
matter,
to
sediments
in
water,
and
to
soil
organic
particles.
Volatilization
from
soil
or
water
to
air
appears
to
occur
relatively
slowly
(
U.
S.
EPA,
1991a).

Very
little
information
is
available
on
degradation
or
transformation
of
HCBD.
Under
aerobic
conditions,
HCBD
in
sewage
contaminated
waters
showed
complete
biodegradation
(
Tabak
et
al.,
1981).
Under
anaerobic
soil
conditions,
biodegradation
will
not
occur
based
on
results
obtained
in
sludge
incubated
under
anaerobic
conditions
(
Johnson
and
Young,
1983).
Estimates
of
the
half­
life
of
HCBD
in
water
range
from
3
to
30
days
in
rivers
and
from
30
to
300
days
in
lakes
and
ground
water
(
Zoeteman
et
al.,
1980).

HCBD
may
readily
partition
from
the
water
into
biological
tissues,
as
suggested
by
its
high
log
octanol:
water
partition
coefficient
(
K
OW
of
4.78).
Laboratory
and
field
studies
have
confirmed
its
bioaccumulation
potential
(
WHO,
1994;
U.
S.
EPA,
1998a).
There
is
no
evidence
that
HCBD
has
biomagnification
potential
(
WHO,
1994).

HCBD
is
not
used
as
a
drinking
water
treatment
chemical,
and
leaching
from
drinking
water
contact
surfaces
is
not
likely.
Therefore,
these
factors
are
not
expected
to
contribute
to
elevated
levels
of
HCBD
in
drinking
water.

9.4
Risk
Reduction
The
third
criterion
asks
if,
in
the
sole
judgment
of
the
Administrator,
regulation
presents
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.
In
evaluating
this
criterion,
EPA
looked
at
the
total
exposed
population,
as
well
as
the
population
exposed
above
the
estimated
HRL.
Estimates
of
the
populations
exposed
and
the
levels
to
which
they
are
exposed
were
derived
from
the
monitoring
results.
These
estimates
are
included
in
Chapter
4
of
this
document
and
summarized
in
Section
9.4.2
below.

In
order
to
evaluate
risk
from
exposure
through
drinking
water,
EPA
considered
the
net
environmental
exposure
in
comparison
to
the
exposure
through
drinking
water.
For
example,
if
exposure
to
a
contaminant
occurs
primarily
through
ambient
air,
regulation
of
emissions
to
air
provides
a
more
meaningful
opportunity
for
EPA
to
reduce
risk
than
does
regulation
of
the
contaminant
in
drinking
water.
In
making
a
preliminary
regulatory
determination,
the
available
information
on
exposure
through
drinking
water
(
Chapter
4)
and
information
on
exposure
through
other
media
(
Chapter
5)
were
used
to
estimate
the
fraction
that
drinking
water
contributes
to
the
total
exposure.
The
EPA
findings
are
discussed
in
Section
9.4.3
below.

In
making
its
preliminary
regulatory
determination,
EPA
also
evaluated
effects
on
potential
sensitive
populations,
including
the
fetus,
infants
and
children.
Sensitive
population
considerations
are
included
in
section
9.4.4.
HCBD
 
February
2003
9­
11
9.4.1
Risk
Reduction
Criterion
Conclusion
Approximately
2
to
5
million
people
are
served
by
systems
with
detections
of
HCBD.
An
estimated
10,000
of
these
individuals
may
be
served
by
systems
with
detections
greater
than
the
HRL,
based
on
Round
2
monitoring
data.
Sensitive
populations
to
HCBD
may
include
people
with
preexisting
kidney
damage
and
infants,
though
it
is
unlikely
for
human
newborns
to
be
acutely
exposed
to
significant
doses
of
HCBD.
When
average
daily
intakes
from
drinking
water
are
compared
with
intakes
from
air,
drinking
water
accounts
for
a
relatively
small
proportion
of
total
HCBD
intake.
Relative
intake
rates
from
food
may
be
higher,
however,
and
intakes
from
soil
are
not
known.
On
the
basis
of
these
observations,
the
impact
of
regulating
HCBD
concentrations
in
drinking
water
on
health
risk
reduction
is
likely
to
be
small.

9.4.2
Exposed
Population
Estimates
National
population
estimates
for
HCBD
exposure
were
derived
using
summary
statistics
for
Round
1
and
Round
2
PWS
cross­
sectional
data
(
see
Table
4­
2
of
this
document)
and
population
data
from
the
Water
Industry
Baseline
Handbook
(
U.
S.
EPA,
2000e).
Summary
data
for
exposed
population
estimates
are
provided
in
Table
9­
2
below.
An
estimated
1.9
to
5
million
people
are
served
by
PWSs
that
have
detected
HCBD.
Of
this
population,
approximately
1.2
million
people
could
be
exposed
at
one­
half
of
the
HRL,
based
on
data
from
Round
1
sampling;
and
about
5
million
people
could
be
exposed
to
over
one­
half
the
HRL,
based
on
Round
2
sampling.
Based
on
the
data
from
Round
1
sampling,
about
781,000
individuals
were
exposed
to
concentrations
at
or
above
the
HRL.
Based
on
Round
2
sampling
results,
an
estimated
10,000
persons
could
be
exposed
at
or
above
the
HRL.
The
Round
2
based
estimate
is
probably
a
better
estimate
of
possible
exposure
since
the
database
is
more
recent,
and
more
representative
of
the
cross­
section
population
served
by
groundwater.

Table
9­
2.
National
Population
Estimates
for
HCBD
Exposure
via
Drinking
Water.

Population
of
Concern
Round
1
Round
2
Served
by
PWS
with
detections
1,909,000
5,027,000
Served
by
PWSs
with
detections
>
(
1/
2
HRL)
1,213,000
4,965,000
Served
by
PWSs
with
detections
>
HRL
781,000
10,000
Source:
Data
taken
from
Table
4­
2
of
this
document.
HRL
=
Health
Reference
Level
9.4.3
Relative
Source
Contribution
Relative
source
contribution
analysis
compares
the
magnitude
of
exposure
expected
via
drinking
water
to
the
magnitude
of
exposure
from
intake
of
HCBD
in
other
media,
such
as
food,
air,
and
soil.
To
perform
this
analysis,
intake
of
HCBD
from
drinking
water
must
be
estimated.
HCBD
 
February
2003
9­
12
Occurrence
data
for
HCBD
in
water
and
other
media
are
presented
in
Chapter
4
and
5
of
this
document.

As
shown
in
Table
4­
2,
the
99th
percentile
concentration
for
all
samples
(
i.
e.,
those
with
detectable
and
nondetectable
levels
of
HCBD)
from
Round
1
and
Round
2
PWS
sampling
is
below
the
MRL.
As
a
convention,
a
value
of
half
the
MRL
is
often
used
as
an
estimate
of
the
concentration
of
a
contaminant
in
samples/
systems
whose
results
are
less
than
the
MRL.
However,
for
Round
1
and
Round
2,
States
have
reported
a
wide
range
of
values
for
the
MRLs
(
See
Section
4.2.1),
and
a
single
estimate
of
the
MRL
for
HCBD
is
unavailable.

As
an
alternative,
the
median
concentration
(
0.3
µ
g/
L)
for
HCBD
in
samples
with
detectable
levels
from
both
rounds
was
used
to
estimate
intake
from
drinking
water.
The
exposure
estimate
for
an
average
individual
is
determined
by
multiplying
the
drinking
water
concentration
by
daily
water
intake
(
2
liters/
day)
and
dividing
by
average
adult
body
weight
(
70
kg),
and
is
estimated
to
be
8.6
×
10­
6
mg/
kg­
day.
For
children,
assuming
a
daily
water
intake
of
1
liter/
day
and
body
weight
of
10
kg,
the
exposure
estimate
is
3.0
×
10­
5
mg/
kg­
day.

The
estimated
average
daily
intakes
of
HCBD
from
drinking
water
(
based
on
median
concentration
of
detected
samples)
and
other
sources
are
shown
in
Table
9­
3.
The
estimated
food:
drinking
water
exposure
ratio
is
0.03
for
an
adult
and
0.02
for
a
child
(
Table
9­
4).
The
estimated
air:
drinking
water
exposure
is
14
for
an
adult
and
21
for
a
child.
Collectively,
these
data
indicate
that
intake
from
drinking
water
is
low
when
compared
to
intake
from
air,
though
not
necessarily
when
compared
to
possible
intake
from
food.

Table
9­
3.
Comparison
of
Average
Daily
Intakes
from
Drinking
Water
and
Other
Media
a.

Medium
Adult
(
ng/
kg­
day)
Child
(
ng/
kg­
day)

Drinking
Water
b
8.6
30
Food
0.15
0.24
Air
120
630
a
See
Chapter
5
for
derivation
of
intakes
from
media
other
than
water
b
Based
on
half
the
median
values
for
detected
hexachlorobutadiene
concentrations
in
Round
1
and
Round
2
Table
9­
4.
Ratios
a
of
Exposures
from
Various
Media
to
Exposures
from
Drinking
Water.

Exposure
Ratio
Adult
Child
Food:
Drinking
Water
0.02
0.008
Air:
Drinking
Water
14
21
a
Calculated
from
estimated
daily
intakes
in
Table
9­
3
HCBD
 
February
2003
9­
13
9.4.4
Sensitive
Populations
The
target
organ
for
HCBD
is
primarily
the
kidney.
Sensitive
populations
to
HCBD
exposure
may
include
people
with
preexisting
kidney
damage.
Though
it
is
unlikely
that
human
newborns
would
be
acutely
exposed
to
significant
doses
of
HCBD,
acute
exposure
for
young
rats
causes
toxicity
at
lower
levels
than
for
adults
(
Hook
et
al.,
1983;
Lock
et
al.,
1984).

Calculation
of
medium­
specific
exposure
ratios
(
Table
9­
4)
indicates
that
HCBD
intake
from
air
is
about
14­
20
fold
greater
than
intake
from
water.
Therefore,
regulation
of
HCBD
in
drinking
water
would
be
unlikely
to
significantly
reduce
the
risk
to
sensitive
populations.

9.5
Regulatory
Determination
Summary
While
there
is
evidence
that
HCBD
may
have
adverse
health
effects
in
humans
at
moderateto
high
doses,
it
is
unlikely
that:
1)
this
contaminant
will
occur
with
a
frequency
or
at
concentrations
that
are
of
public
health
concern;
or
2)
regulation
of
this
contaminant
represents
a
meaningful
basis
for
health
risk
reduction
in
persons
served
by
public
water
systems.
For
these
reasons,
EPA
does
not
plan
to
regulate
HCBD
with
a
NPDWR..
HCBD
 
February
2003
10­
1
10.0
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M.
Abdo
and
M.
R.
Elwell.
1989.
Subchronic
toxicology
studies
of
hexachloro­
1,3­
butadiene
(
HCBD)
in
B6C3F
1
mice
by
dietary
incorporation.
J.
Env.
Path.
Tox.
&
Onc.
9:
323­
332.

Yip,
G.
1976.
Survey
of
hexachloro­
1,3,­
butadiene
in
fish,
eggs,
milk,
and
vegetables.
J.
Assoc.
Off.
Anal.
Chem.
59:
559­
561.

Yurawecz,
M.
P.,
P.
A.
Dreifuss
and
L.
R.
Kamps.
1976.
Determination
of
hexachloro­
1,3­
butadiene
in
spinach,
eggs,
fish,
and
milk
by
electron
capture
gas­
liquid
chromatography.
J.
Assoc.
Off.
Anal.
Chem.
59:
552­
558.

Zoeteman,
B.
C.
J.,
K.
Harmsen,
J.
B.
H.
J.
Linders,
et
al.
1980.
Persistent
organic
pollutants
in
river
water
and
groundwater
of
the
Netherlands.
Chemosphere
9:
231­
249.
HCBD
 
February
2003
A­
1
APPENDIX
A:
Abbreviations
and
Acronyms
ATSDR
­
Agency
for
Toxic
Substances
and
Disease
Registry
CAS
­
Chemical
Abstract
Service
CCL
­
Contaminant
Candidate
List
CERCLA
­
Comprehensive
Environmental
Response,
Compensation
&
Liability
Act
CMR
­
Chemical
Monitoring
Reform
CWS
­
Community
Water
System
DWEL
­
Drinking
Water
Equivalent
Level
EPA
­
Environmental
Protection
Agency
EPCRA
­
Emergency
Planning
and
Community
Right­
to­
Know
Act
FDA
­
Food
and
Drug
Administration
GW
­
ground
water
HRL
­
Health
Reference
Level
IRIS
­
Integrated
Risk
Information
System
MCL
­
Maximum
Contaminant
Level
MRL
­
Minimum
Reporting
Level
NAWQA
­
National
Water
Quality
Assessment
Program
NCOD
­
National
Drinking
Water
Contaminant
Occurrence
Database
NIOSH
­
National
Institute
for
Occupational
Safety
and
Health
NPDWR
­
National
Primary
Drinking
Water
Regulation
NPL
­
National
Priorities
List
NTIS
­
National
Technical
Information
Service
NTNCWS
­
Non­
Transient
Non­
Community
Water
System
ppm
­
part
per
million
PWS
­
Public
Water
System
SARA
Title
III
­
Superfund
Amendments
and
Reauthorization
Act
SDWA
­
Safe
Drinking
Water
Act
SDWIS
­
Safe
Drinking
Water
Information
System
SDWIS
FED
­
the
Federal
Safe
Drinking
Water
Information
System
SOC
­
synthetic
organic
compound
STORET
­
Storage
and
Retrieval
System
SW
­
surface
water
TRI
­
Toxic
Release
Inventory
UCM
­
Unregulated
Contaminant
Monitoring
UCMR
­
Unregulated
Contaminant
Monitoring
Regulation/
Rule
URCIS
­
Unregulated
Contaminant
Monitoring
Information
System
U.
S.
EPA
­
United
States
Environmental
Protection
Agency
USGS
­
United
States
Geological
Survey
VOC
­
volatile
organic
compound
:
g/
L
­
micrograms
per
liter
mg/
L
­
milligrams
per
liter
>
MCL
­
percentage
of
systems
with
exceedances
>
MRL
­
percentage
of
systems
with
detections
HCBD
 
February
2003
B­
1
APPENDIX
B:
Round
1
and
Round
2
Occurrence
Data
Tables
for
Hexachlorobutadiene
Hexachlorobutadiene
Occurrence
in
Public
Water
Systems
in
Round
1,
UCM
(
1987)
results
STATE
TOTAL
UNIQUE
PWS
#
GW
PWS
#
SW
PWS
%
PWS
with
detections
%
GW
PWS
with
detections
%
SW
PWS
with
detections
%
PWS
>
HRL
%
GW
PWS
>
HRL
%
SW
PWS
>
HRL
99%
VALUE
(
µ
:
g/
L)

AK
665
540
130
1.50%
1.48%
1.54%
0.00%
0.00%
0.00%
<
0.00
AL
131
93
42
3.05%
4.30%
0.00%
1.53%
2.15%
0.00%
0.50
AR
AZ
448
407
47
0.89%
0.74%
2.13%
0.22%
0.00%
2.13%
<
2.00
CA
585
571
21
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
6.00
CO
6
3
4
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.64
DC
1
0
1
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
DE
10
8
2
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
FL
112
7
105
5.36%
0.00%
5.71%
5.36%
0.00%
5.71%
5.00
GA
HI
127
112
16
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.30
IA
IL
213
149
64
0.47%
0.67%
0.00%
0.00%
0.00%
0.00%
<
2.00
IN
357
321
37
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
2.00
KY
524
291
233
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
1.00
LA
13
9
4
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
MA
MD
983
936
50
0.10%
0.11%
0.00%
0.00%
0.00%
0.00%
<
0.50
MI
MN
1,553
1,529
28
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
MO
85
71
14
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
20.00
MS
MT
NC
297
254
44
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
NE
NH
NJ
801
790
11
0.75%
0.76%
0.00%
0.25%
0.25%
0.00%
<
1.20
NM
590
555
35
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
1.00
NV
8
7
2
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.20
NY
356
252
123
0.28%
0.40%
0.00%
0.28%
0.40%
0.00%
<
5.00
OH
2,655
2,493
166
0.11%
0.12%
0.00%
0.08%
0.08%
0.00%
<
2.00
SD
335
306
29
0.30%
0.33%
0.00%
0.00%
0.00%
0.00%
<
0.50
TN
303
156
147
0.33%
0.64%
0.00%
0.33%
0.64%
0.00%
<
0.50
TX
2
2
0
100.00%
100.00%
0.00%
100.00%
100.00%
0.00%
8.00
UT
411
391
34
1.22%
1.02%
2.94%
0.00%
0.00%
0.00%
<
5.00
VI
3
0
3
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
1.00
VT
WA
992
937
77
0.10%
0.11%
0.00%
0.00%
0.00%
0.00%
<
0.50
WV
57
26
31
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
4.00
WY
145
116
38
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
2.00
TOTAL
12,768
11,332
1,538
0.36%
0.32%
0.65%
0.12%
0.07%
0.46%
<
5.00
24
STATE
S
12,284
10,980
1,385
0.35%
0.30%
0.72%
0.11%
0.06%
0.51%
<
5.00
PWS
=
Public
Water
Systems;
GW
=
Ground
Water;
SW
=
Surface
Water;
MRL
=
Minimum
Reporting
Limit
(
for
laboratory
analyses);
Health
Reference
Level
=
Health
Reference
Level,
an
estimated
health
effect
level
used
for
preliminary
assessment
for
this
review
The
Health
Reference
Level
used
for
hexachlorobutadiene
is
0.9
µ
:
g/
L.
This
is
a
draft
value
for
working
review
only.
Total
Number
of
PWSs
=
the
total
number
of
public
water
systems
with
records
for
hexachlorobutadiene
%
PWS
with
detections,
>
½
Health
Reference
Level,
>
Health
Reference
Level
=
percent
of
the
total
number
of
public
water
systems
with
at
least
one
analytical
result
that
exceeded
the
MRL,
½
Health
Reference
Level,
Health
Reference
Level,
respectively
99th
Percentile
Concentration
=
the
concentration
value
of
the
99th
percentile
of
all
analytical
results
(
in
µ
:
g/
L)
Median
Concentration
of
Detections
=
the
median
analytical
value
of
all
the
detections
(
analytical
results
greater
than
the
MRL)
(
in
µ
:
g/
L)
The
highlighted
states
are
part
of
the
URCIS
(
Round
1)
24
State
Cross­
Section.
HCBD
 
February
2003
B­
2
Hexachlorobutadiene
Occurrence
in
Public
Water
Systems
in
Round
2,
UCM
(
1993)
results
STATE
TOTAL
UNIQUE
PWS
#
GW
PWS
#
SW
PWS
%
PWS
with
detections
%
GW
PWS
with
detections
%
SW
PWS
with
detections
%
PWS
>
HRL
%
GW
PWS
>
HRL
%
SW
PWS
>
HRL
99%
VALUE
(
µ
:
g/
L)

Tribes
(
06)
22
21
1
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
50.00
AK
625
481
144
3.36%
2.70%
5.56%
0.00%
0.00%
0.00%
<
0.00
AL
AR
407
319
88
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.10
AZ
68
60
8
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
1.00
CA
14
11
3
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
CO
831
619
212
0.24%
0.00%
0.94%
0.00%
0.00%
0.00%
<
0.00
CT
84
43
41
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
IN
117
107
10
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
2.00
KY
121
50
71
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
2.50
LA
1,310
1,241
69
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
MA
418
344
74
0.24%
0.00%
1.35%
0.24%
0.29%
0.00%
<
0.50
MD
976
920
56
0.20%
0.11%
1.79%
0.00%
0.00%
0.00%
<
0.50
ME
744
676
68
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
MI
2,739
2,647
92
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
MN
1,558
1,528
30
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
MO
1,412
1,297
115
0.07%
0.08%
0.00%
0.00%
0.00%
0.00%
<
1.00
MS
1
1
100.00%
100.00%
0.00%
0.00%
0.60
NC
1,775
1,585
190
0.51%
0.44%
1.05%
0.00%
0.00%
0.00%
<
0.00
ND
296
258
38
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
NH
NJ
7
7
0.00%
0.00%
0.00%
0.00%
<
1.00
NM
720
693
27
0.14%
0.14%
0.00%
0.00%
0.00%
0.00%
<
1.00
OH
2,232
2,050
182
0.04%
0.05%
0.00%
0.04%
0.00%
0.55%
<
0.50
OK
790
541
249
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
OR
17
15
2
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
PA
RI
115
103
12
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
1.00
SC
237
216
21
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
SD
27
19
8
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
TN
TX
4,412
3,825
587
0.07%
0.08%
0.00%
0.05%
0.00%
0.34%
1.00
VT
1
1
0.00%
0.00%
0.00%
0.00%
<
0.50
WA
2,548
2,429
119
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
WI
191
188
3
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.30
TOTAL
24,815
22,294
2,521
0.17%
0.13%
0.56%
0.02%
0.00%
0.12%
<
1.00
20
STATES
22,736
20,380
2,356
0.18%
0.13%
0.59%
0.02%
0.00%
0.13%
<
1.00
PWS
=
Public
Water
Systems;
GW
=
Ground
Water;
SW
=
Surface
Water;
MRL
=
Minimum
Reporting
Limit
(
for
laboratory
analyses);
Health
Reference
Level
=
Health
Reference
Level,
an
estimated
health
effect
level
used
for
preliminary
assessment
for
this
review
The
Health
Reference
Level
used
for
hexachlorobutadiene
is
0.9
:
g/
L.
This
is
a
draft
value
for
working
review
only.
Total
Number
of
PWSs
=
the
total
number
of
public
water
systems
with
records
for
hexachlorobutadiene
%
PWS
with
detections,
>
½
Health
Reference
Level,
>
Health
Reference
Level
=
percent
of
the
total
number
of
public
water
systems
with
at
least
one
analytical
result
that
exceeded
the
MRL,
½
Health
Reference
Level,
Health
Reference
Level,
respectively
99th
Percentile
Concentration
=
the
concentration
value
of
the
99th
percentile
of
all
analytical
results
(
in
:
g/
L)
Median
Concentration
of
Detections
=
the
median
analytical
value
of
all
the
detections
(
analytical
results
greater
than
the
MRL)
(
in
µ
g/
L)
The
highlighted
States
are
part
of
the
SDWIS/
FED
(
Round
2)
20
State
Cross­
Section.
