Document ID: EPA-HQ-OPP-2004-0402-0031
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2005-03-18T05:00Z

Hexachlorobenzene
(
HCB)
As
a
Contaminant
of
Pentachlorophenol
Ecological
Hazard
and
Risk
Assessment
for
the
Pentachlorophenol
Reregistration
Eligibiltiy
Decision
(
RED)
Document
3/
4/
05
i
TABLE
OF
CONTENTS
Page
No.

LIST
OF
TABLES
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
ii
LIST
OF
APPENDICES
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
ii
I.
EXECUTIVE
SUMMARY
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
1
II.
BACKGROUND/
INTRODUCTION
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
2
III.
TOXICITY
OF
HEXACHLOROBENZENE
(
HCB)
TO
AQUATIC
ORGANISMS
.
.
.
3
A.
Fish,
Aquatic
Invertebrates,
and
Plants
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
3
B.
Sediment
Toxicity
of
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
6
C.
Bioaccumulation
of
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
6
D.
Conclusions
Concerning
Aquatic
Toxicity
of
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
8
IV.
TOXICITY
OF
HCB
TO
TERRESTRIAL
ORGANISMS
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
8
A.
Toxicity
Assessment
for
Invertebrates
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
9
B.
Toxicity
Assessment
for
Birds
and
Bird
Eggs
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
9
C.
Toxicity
Assessment
for
Mammals
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
11
D
Conclusions
Concerning
Terrestrial
Toxicity
of
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
12
V.
ECOLOGICAL
RISK
ASSESSMENT
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
13
A.
Risk
to
Terrestrial
Organisms
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
15
B.
Risk
to
Aquatic
Organisms
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
17
C.
Risk
to
Plants
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
19
VI.
ENDANGERED
SPECIES
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
19
VII.
UNCERTAINTIES
IN
THIS
ASSESSMENT
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
20
VIII.
REFERENCES
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
21
ii
LIST
OF
TABLES
Page
No.

Table
1.
Selected
Aquatic
Toxicity
Values
for
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
4
Table
2.
Whole
Sediment
Acute
Toxicity
Values
for
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
6
Table
3.
Selected
Avian
Toxicity
Values
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
9
Table
4.
Selected
Mammalian
Toxicity
Values
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
11
Table
5.
Summary
of
HCB
Concentrations
in
Air,
Water,
Sediments,
and
Soils,
as
Calculated
in
U.
S.
EPA
(
2002)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
16
Table
6.
HCB
Acute
and
Chronic
RQs
for
Freshwater
Fish
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
17
Table
7.
HCB
Acute
and
Chronic
RQs
for
Freshwater
Invertebrates
.
.
.
.
.
.
.
.
.
.
.
.
.
17
Table
8.
HCB
Acute
RQ
for
Estuarine/
Marine
Fish
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
18
Table
9.
HCB
Acute
RQ
for
Aquatic
Plants
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
19
LIST
OF
APPENDICES
Appendix
1.
Toxicity
Data
for
HCB
(
EPA,
1985
 
Health
Assessment
for
Chlorinated
Benzenes)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
29
Appendix
2.
Toxicity
Values
for
Aquatic
Animals
and
Plants
(
EPA,
1988)
.
.
.
.
.
.
.
.
.
.
.
30
Appendix
3.
Aquatic
Toxicity
Data
from
ECOTOX
Database
(
EPA,
2001)
.
.
.
.
.
.
.
.
.
.
33
Appendix
4.
Additional
Aquatic
Toxicity
Data
­
Miscellaneous
Sources
.
.
.
.
.
.
.
.
.
.
.
.
.
35
Appendix
5.
Terrestrial
Toxicity
Data
(
Invertebrates,
Birds,
Mammals)
 
Ecotox
Database,
EPA
2001
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
36
Appendix
6.
Terrestrial
Toxicity
Data
­
EPA
1985,
Health
Assessment
for
Chlorobenzenes
(
Mammals
Only)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
38
Appendix
7.
Terrestrial
Toxicity
Data
(
Mammals)
­
ATSDR,
2000
 
Toxicity
Profile
for
HCB
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
40
Appendix
8.
Terrestrial
Toxicity
Data
 
Miscellaneous
Sources
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
42
Appendix
9.
Terrestrial
Exposure
Assessment
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
43
1
I.
EXECUTIVE
SUMMARY
Hexachlorobenzene
(
HCB)
is
formed
as
a
by­
product
during
the
manufacture
of
pentachlorophenol
and
other
chlorine­
containing
compounds.
HCB
released
to
the
environment
tends
to
persist
in
soil
and
surface
water
for
a
long
time,
with
recorded
half­
life
values
ranging
from
3­
6
years
and
longer.
This
compound
also
readily
bioaccumulates
in
both
aquatic
and
terrestrial
organisms.

Data
on
the
ecological
effects
of
HCB
are
relatively
limited.
There
are
no
Guideline
studies
for
HCB,
since
it
is
not
registered
as
a
pesticide.
The
majority
of
data
presented
in
this
report
are
based
on
information
compiled
in
EPA
technical
reports
(
U.
S.
EPA,
1985;
U.
S
EPA,
1988;
U.
S.
EPA,
2001)
and
the
Agency
for
Toxic
Substances
and
Disease
Registry's
Toxicological
Profiles
for
HCB
(
ATSDR,
2000).

Reviewed
scientific
literature
indicates
that
HCB
has
a
limited
potential
to
adversely
affect
aquatic
organisms
in
the
short­
term,
primarily
due
to
its
very
low
solubility
in
surface
water.
The
majority
of
the
toxicological
data
for
fish,
benthic
invertebrates
and
aquatic
plants
did
not
indicate
any
lethal
or
significant
sublethal
effects
from
acute
or
chronic
exposure
to
HCB
at
its
saturation
concentration
of
6

g/
L.
Some
studies
indicate
that
HCB
concentrations
in
the
tissues
of
aquatic
organisms
appear
to
equilibrate
very
slowly
with
concentrations
in
the
water.
As
a
result,
the
reviewed
chronic
toxicity
tests
for
fish
species
(
e.
g.,
rainbow
trout
and
fathead
minnows)
may
not
have
been
of
sufficient
duration
to
allow
for
the
full
equilibration
of
HCB
in
fish
tissue
with
surface
water
concentrations.
Also,
due
to
the
tendency
of
HCB
to
bioaccumulate
in
the
aquatic
foodweb,
there
is
the
potential
for
adverse
effects
to
higher­
trophic
level
organisms
from
exposure
to
HCB
in
their
diet.

HCB
is
slightly
to
moderately
toxic
to
terrestrial
birds
and
mammals.
LD50
values
for
evaluated
avian
species
ranged
from
568
mg/
kg
for
Japanese
quail
(
exposed
to
HCB
over
a
5­
day
period)
to
greater
than
5000
mg/
kg
for
mallard
ducks
(
exposed
to
HCB
for
8
days).
LD50
values
for
mammals
ranged
from
710
mg/
kg
for
deer
mice
(
exposed
to
HCB
over
a
three
day
period)
to
4,000
mg/
kg
for
mice
(
species
and
exposure
duration
not
reported).

For
chronic
exposure,
the
lowest
reported
NOAELs
for
birds
and
mammals
were
1
mg/
kg
in
Japanese
quail
(
over
a
90­
day
period)
and
1.6
mg/
kg­
BW
per
day
in
rats
(
over
a
duration
greater
than
247
days),
respectively.
Sublethal
effects
reported
for
birds
and
mammals
were
often
similar
and
included
hepatic
toxicity,
porphyria,
reproductive
effects
and
reduced
birth
weight.

Hexachlorobenzene
accumulated
in
birds
is
excreted
into
the
egg,
which
results
in
uptake
by
the
embryo.
HCB
concentrations
measured
in
the
eggs
of
sea
birds
and
raptors
from
a
number
of
locations
from
around
the
world
approach
those
associated
with
reduced
embryo
weights
in
herring
gulls
(
1.5
mg/
kg),
suggesting
that
HCB
has
the
potential
to
harm
embryos
of
avian
species.
For
mammals,
a
sensitive
endpoint
for
chronic
HCB
exposure
is
the
reduction
of
birth
weight
and
increased
mortality
in
mink
offspring
exposed
to
1
ppm
HCB
(
0.16
mg/
kg
BW­
day)
for
47
weeks.
This
observation
is
ecologically
significant
because
field
studies
have
observed
HCB
concentrations
in
fish
tissue
at
a
number
of
sites
worldwide
that
are
within
an
order
of
2
magnitude
of
the
dietary
toxicity
level
of
1
ppm.
This
suggests
that
HCB
has
the
potential
to
cause
adverse
effects
in
mink
and
perhaps
other
fish­
eating
mammals,
especially
given
HCB's
tendency
to
bioaccumulate.

Reviewed
literature
indicate
that
HCB
does
have
the
potential
to
adversely
affect
both
aquatic
and
terrestrial
organisms,
due
to
its
persistence
in
the
environment
and
its
ability
to
readily
accumulate
in
the
aquatic
and
terrestrial
foodwebs.

II.
BACKGROUND/
INTRODUCTION
HCB
is
a
very
stable
chlorinated
aromatic
compound
that
was
commonly
used
as
a
pesticide
until
1965.
Currently,
there
are
no
commercial
uses
of
the
substance
in
the
United
States.
HCB
may
be
formed
as
a
byproduct
during
the
manufacture
of
chemicals
used
as
solvents,
pesticides
and
other
chlorine­
containing
compounds.
Small
amounts
of
this
compound
can
also
be
produced
during
combustion
processes
such
as
burning
of
city
wastes.

HCB
is
widely
distributed
throughout
the
global
ecosystem
because
if
its
mobility
and
resistance
to
degradation.
It
has
been
detected
in
all
environmental
media
and
in
numerous
types
of
living
organisms
including
insects,
aquatic
biota,
birds
and
mammals.
HCB
has
also
been
shown
to
bioaccumulate
in
both
aquatic
and
terrestrial
organisms.

This
report
presents
a
compilation
and
evaluation
of
the
available
ecological
effects
data
for
HCB
on
aquatic
life
(
e.
g.,
fish,
arthropods,
etc.),
terrestrial
invertebrates,
birds
and
mammals
(
e.
g.,
rats,
monkeys).
The
majority
of
the
data
presented
in
this
report
are
based
on
compilations
of
data
in
U.
S.
EPA
(
1985),
U.
S.
EPA
(
1988),
the
EPA
Ecotox
database
(
U.
S.
EPA,
2001)
and
ATSDR
(
2000).
Toxicity
endpoint
data
have
been
provided
in
tabular
format
and
a
brief
discussion
of
certain
studies
is
also
provided.
In
some
cases,
data
from
the
same
study
are
presented
in
multiple
tables.
The
reader
is
referred
to
U.
S.
EPA
(
1985;
1988;
2001)
and
ATSDR
(
2000)
for
a
more
detailed
discussion
of
the
toxicity
studies
used
therein.
3
III.
TOXICITY
OF
HEXACHLOROBENZENE
TO
AQUATIC
ORGANISMS
A.
Fish,
Aquatic
Invertebrates,
and
Plants
While
there
are
no
EPA
Guideline
studies
available
for
HCB,
literature
data
provide
some
studies
with
comparable
endpoints.
The
primary
sources
of
aquatic
toxicity
data
used
for
this
report
are
described
below.

EPA's
final
report
entitled
"
Health
Assessment
Document
for
Chlorinated
Benzenes"
published
in
1985
summarized
the
current
knowledge
of
chlorinated
benzenes
in
the
environment
(
including
fate,
transport
and
distribution)
and
current
literature
regarding
toxicological
effects
of
chlorinated
benzenes
on
laboratory
animals
and
humans.
The
report
included
an
ecological
section
focused
on
the
effects
of
chlorinated
benzenes
on
aquatic
organisms
including
fish,
crustaceans
and
plants.
Laboratory
test
concentrations
and
effects
data
reported
in
U.
S.
EPA
(
1985)
are
summarized
for
HCB
in
Appendix
1.

EPA's
Draft
1988
Report
entitled
"
Ambient
Aquatic
Life
Water
Quality
Criteria
for
Hexachlorobenzene"
developed
by
the
Office
of
Research
and
Development,
Environmental
Research
Laboratory
evaluated
the
current
scientific
literature
regarding
acute
and
chronic
toxicity
data
for
HCB
on
aquatic
animals
and
plants
to
develop
a
national
water
quality
criteria
for
hexachlorobenzene,
specifically
for
the
protection
of
aquatic
organisms.
Acute
and
chronic
toxicity
data
reported
in
U.
S.
EPA
(
1988)
are
summarized
in
Appendix
2.

EPA
also
maintains
the
ECOTOX
database,
which
represents
an
integration
of
the
AQUIRE,
PHYTOTOX
and
TERRETOX
databases
that
contain
toxicological
data
for
aquatic
life,
terrestrial
plants
and
terrestrial
wildlife,
respectively.
Summary
toxicity
data
from
ECOTOX
regarding
the
effects
of
HCB
on
aquatic
organisms
are
presented
in
Appendix
3.
The
data
included
in
the
appendix
are
only
those
data
where
a
toxicity
endpoint
was
provided.
This
database
includes
toxicity
studies
conducted
through
2000,
and
therefore
has
more
recent
toxicity
data
than
U.
S.
EPA
1985
and
1988.

In
addition,
aquatic
toxicity
data
for
HCB
collected
from
other
sources,
such
as
scientific
journals
and
technical
reports,
were
also
reviewed,
and
those
data
that
were
not
already
included
in
Appendices
1­
3
are
provided
in
Appendix
4.

Most
of
the
reviewed
aquatic
toxicity
studies
for
HCB
have
been
conducted
as
acute
or
subchronic
exposures.
Selected
acute
and
chronic
endpoint
values
from
Appendices
1­
4
have
been
summarized
and
provided
in
Table
1.
4
Table
1.
Selected
Aquatic
Toxicity
Values
for
Hexachlorobenzene
Species
Test
Type/
Duration
Endpoint
Concentration/
Effect
Source
Acute
and
Subchronic
Toxicity
Channel
catfish
(
Ictalurus
punctatus)
Static
 
24
hrs.
LC50
 
6400

g/
L
Mayer
and
Ellersieck,
1986
Bluegill
(
Lepomis
macrochirus)
Static
 
duration
not
reported
LC50
 
12,000

g/
L
Johnson
and
Finley,
1980
Largemouth
bass
(
Micropterus
salmoides)
Static
 
96
hrs.
LC50
 
12,000

g/
L
Johnson
and
Finley,
1980
Coho
salmon
Static
 
96
hrs.
LC50
 
50,000

g/
L
Johnson
and
Finley,
1980
Flounder
(
Platichthys
flesus)
96
hrs.
LC50
 
200

g/
L
Furay
and
Smith,
1995
Sole
(
Solea
solea)
96
hrs.
LC50
 
140

g/
L
Furay
and
Smith,
1995
Largemouth
bass
(
Micropterus
salmoides)
10
days
3.5

g/
L
(
liver
and
kidney
damage)
Laseter
et
al.,
1976
Diatom
(
Cyclotella
meneghiniana)
Static
­
48
hrs.
EC50
­
2

g/
L
(
genetic
effect)
Figueroa
et
al.,
1991
Freshwater
green
algae
(
Selanastrum
capricornutum)
Static
­
96
hours
EC50
>
30

g/
L
Calamari
et
al.,
1983
Chronic
Toxicity
Water
flea
(
Daphnia
magna)
14
days
EC50
 
16

g/
L
(
50percent
reduction
in
fertility)
Calamari
et
al.,
1983
Cladoceran,
(
Ceriodaphnia
dubia)
Life
cycle/
partial
life
cycle
 
7
days
NOEC
>
7

g/
L
Spehar,
1986
Rainbow
trout
(
Salmo
gairdneri)
Early
life
stage
 
90
days
NOEC
>
3.68

g/
L
Spehar,
1986
Fathead
minnow
(
Pimephales
promelas)
Early
life
stage
 
32
days
NOEC
>
4.8f

g/
L
Ahmad
et
al.,
1984:
Carlson
and
Kosian,
1987
Snail
(
Lymnaea
palustris)
10­
12
weeks
0.5

g/
L,
5

g/
L
(
significantly
reduced
growth
rate)
Baturo
et
al.,
1995
Crab
(
Portunus
pelagicus)
6
weeks
5

g/
L
(
reduction
in
carapace
growth
rate)
Mortimer
and
Connell,
1995
HCB
has
a
very
low
solubility
in
water
(
6

g/
L)
(
ATSDR,
2000).
As
a
result,
the
toxicity
of
HCB
to
aquatic
organisms
is
generally
considered
a
function
of
its
solubility.
Literature
studies
indicate
that
the
only
experimentally
derived
acute
toxicity
values
for
HCB
were
at
concentrations
that
exceeded
the
solubility
limit
for
this
compound,
in
some
cases
by
several
orders
of
magnitude.
For
example,
Mayer
and
Ellersieck
(
1986)
determined
that
the
LC50
for
channel
catfish
exposed
to
HCB
for
24
hours
ranged
between
6,400
and
16,200

g/
L.
Similarly,
Johnson
and
Findley
(
1980)
reported
that
largemouth
bass
and
coho
salmon
exposed
5
to
HCB
for
96
hours
had
LC50
values
of
12,000

g/
L
and
50,000

g/
L,
respectively
(
Table
1).
They
noted
that
the
HCB
concentrations
for
the
experiments
were
obtained
using
a
cosolvent
to
increase
HCB's
solubility.

Acute
toxicity
data
compiled
in
Appendices
1­
4
appear
to
support
the
conclusion
that
HCB
concentrations
that
are
within
an
order
of
magnitude
of
the
saturation
concentration
of
6

g/
L
do
not
appear
to
result
in
mortality
to
aquatic
organisms.
The
lack
of
mortality
is
evident
in
the
Appendices
where
the
LC50s
are
consistently
higher
(
e.
g.,
">")
than
the
maximum
concentration
of
HCB
evaluated
in
the
study,
with
study
concentrations
ranging
between
3.7

g/
L
and
100

g/
L.
The
lowest
identified
LC50
concentration
for
acute
toxicity
at
96
hours
exposure
was
approximately
200

g/
L
for
flounder
and
140

g/
L
for
sole
(
Furay
and
Smith,
1995)
as
shown
in
Table
1.
These
data
indicate
that
ambient
concentrations
of
HCB
near
the
solubility
limit
are
not
anticipated
to
result
in
acute
toxicity
to
the
test
species.

Some
studies
of
acute
exposures
were
continued
longer
than
96
hours
with
no
resultant
mortalities.
For
example,
adult
crayfish
(
Procambarus
clarki)
were
unaffected
over
a
period
of
10
days
at
a
mean
HCB
concentration
of
27.3

g/
L,
and
largemouth
bass
(
Micropterus
salmoides)
were
unaffected
over
a
period
of
10
days
at
25.8

g/
L
(
Laseter
et
al.,
1976
and
Laska
et
al.,
1978)
(
Appendix
2).
Two
studies
did,
however,
indicate
sublethal
effects
for
acute
and
subchronic
exposure.
As
shown
in
Table
1,
Laseter
et
al.
(
1978)
found
that
largemouth
bass
exposed
to
a
HCB
concentration
of
3.5

g/
L
for
10
days
did
exhibit
some
liver
and
kidney
damage.
Also,
Figueroa
et
al.
(
1991)
found
that
diatoms
exposed
to
HCB
for
48
hours
had
an
EC50
of
2

g/
L
for
genetic
effects
(
Table
1).

Chronic
aquatic
toxicity
of
HCB
has
also
been
investigated.
Rainbow
trout
(
Salmo
gairdneri)
were
exposed
to
HCB
in
a
90­
day
early
life­
stage
test
(
Spehar,
1986).
As
shown
in
Table
1,
no
adverse
effects
on
hatching,
survival,
or
growth
were
observed
at
the
highest
tested
concentration
of
3.68

g/
L.
Similarly,
juvenile
fathead
minnows
(
Pimephales
promelas)
were
not
affected
at
HCB
exposures
up
to
4.8

g/
L
in
a
32­
day
early
life
stage
test
(
Ahmad
et
al.,
1984;
Carlson
and
Kosian,
1987)
(
Table
1).
Analysis
determined
that
survival
at
hatch,
survival
at
32
days,
and
wet
weight
for
the
fathead
minnows
exposed
to
HCB
were
equivalent
to
the
survival
and
wet
weight
for
the
control
individuals.
In
a
7­
day
life­
cycle
test
with
the
Cladoceran
(
Ceriodaphnia
dubia),
exposure
to
HCB
did
not
result
in
any
measurable
effect
upon
survival
or
reproduction
at
concentrations
as
high
as
7.0

g/
L
(
Spehar,
1986)
(
Table
1).
These
studies
indicate
that
a
HCB
concentration
of
3.68

g/
L
does
not
appear
to
cause
chronic
toxicity
in
any
of
the
tested
species.
However,
based
on
review
of
the
literature,
two
studies
did
indicate
adverse
effects
on
aquatic
organisms
from
chronic
exposure
to
HCB
at
concentrations
below
the
solubility
limit.
Baturo
et
al.
(
1995)
noted
that
snails
(
Lymnaea
palustris)
exposed
to
HCB
concentrations
of
0.5

g/
L
and
5

g/
L
for
a
duration
of
10­
12
weeks
had
a
significantly
reduced
growth
rate
compared
with
snails
in
a
control
group
(
although
the
percentage
difference
in
growth
between
the
two
groups
was
not
reported).
Similarly,
Mortimer
and
Connell
(
1995)
noted
that
crab
carapace
growth
was
reduced
for
juveniles
exposed
to
HCB
concentrations
of
5

g/
L
for
a
duration
of
6
weeks
compared
with
individuals
in
a
control
group,
however,
they
did
not
estimate
an
EC10
or
EC50
because
a
significant
departure
from
the
control
group
was
only
observed
at
one
concentration.
In
6
addition,
Calamari
et
al.
(
1983)
determined
that
water
fleas
(
Daphnia
magna)
exposed
to
HCB
for
14
days
had
an
EC50
of
16

g/
L
for
a
reduction
in
fertility,
although
it
should
be
noted
that
this
concentration
is
higher
than
the
solubility
limit.

B.
Sediment
Toxicity
of
Hexacholorobenzene:

Although
toxicity
of
HCB
in
surface
water
has
been
extensively
studied,
the
toxicity
of
HCB
absorbed
to
sediment
is
less
clearly
understood
(
Barber
et
al.,
1997).
Organisms
can
be
exposed
to
HCB
through
sediment
ingestion,
as
well
as
through
direct
contact
with
surface
water
(
Kolok
et
al.,
1996).
Barber
et
al.
(
1997)
evaluated
potential
effects
to
the
amphipod
(
Hyalella
azteca)
and
midge
(
Chironomus
tentans)
from
exposure
to
HCB
in
sediment.
The
study
results
indicated
that
there
were
no
adverse
effects
to
survival
or
growth
of
the
test
species
at
the
highest
sediment
concentration
of
HCB
tested
(
42
mg/
kg).
The
authors
noted
that
the
lack
of
toxicity
in
this
study
even
at
the
maximum
HCB
concentration
is
consistent
with
available
literature
on
aquatic
toxicity
of
this
chemical.
Fuchsman
et
al.
(
1998)
also
found
that
no
significant
toxicity
occurred
for
the
test
species
(
amphipod
(
Hyalella
azteca)
and
midge
(
Chironomus
tentans))
at
a
HCB
sediment
concentration
of
60
mg/
kg,
adding
to
the
weight
of
evidence
of
limited
potential
for
HCB­
related
sediment
toxicity
to
benthic
invertebrates.
They
concluded
that
sediment
quality
guidelines
such
as
the
Ontario
Ministry
of
the
Environment's
lowest
effect
level
for
HCB
(
0.02
mg/
kg)
is
likely
to
be
overly
conservative
with
respect
to
effects
on
benthic
macroinvertebrates.
A
summary
of
sediment
toxicity
endpoints
is
provided
in
the
table
below.

Table
2.
Whole
Sediment
Acute
Toxicity
of
HCB
to
Aquatic
Invertebrates
Substance/
%
Active
Ingredient
(
AI)
Organism
Endpoints/
Results
(
mg/
kg)
Comments
Reference
HCB
Amphipod
(
Hyalella
azteca)
NOEC
=
42
14­
day
test
duration
(
spiked
sediment)
Barber
et
al.,
1997
HCB
Midge
(
Chironomus
tentans)
NOEC
=
42
14­
day
test
duration
(
spiked
sediment)
Barber
et
al.,
1997
HCB
Freshwater
amphipod
(
Hyalella
azteca)
NOEC
=
60
10­
day
test
duration
(
spiked
sediment)
Fuchsman
et
al.,
1998
HCB
Midge
(
Chironomus
tentans)
NOEC
=
60
10­
day
test
duration
(
spiked
sediment)
Fuchsman
et
al.,
1998
HCB
Estuarine
amphipod
(
Leptochirus
plumulosus)
NOEC
=
60
10­
day
test
duration
(
spiked
sediment)
Fuchsman
et
al.,
1998
C.
Bioaccumulation
of
Hexachlorobenzene:

Although
the
endpoint
toxicity
data
for
HCB
in
abiotic
media
appear
somewhat
limited,
there
is
some
evidence
that
bioaccumulation
in
lower
trophic
level
species
may
result
in
effects
to
higher
trophic
level
species.
For
example,
Munoz
et
al.
(
1996)
evaluated
the
effects
of
HCB
7
accumulation
in
algal
tissue
on
the
water
flea
(
Daphnia
magna).
They
noted
that
uptake
of
HCB
in
algae
(
Chlorella
sp.)
was
rapid
with
an
equilibrium
time
of
2­
4
hours.
The
algae
then
had
a
HCB
concentration
of
approximately
4.4
mg/
kg
dry
weight.
Adult
Daphnia
fed
the
contaminated
algae
for
six
days
had
a
hexachlorobenzene
concentration
of
1.7
mg/
kg
dry
weight.
Although
there
was
no
mortality
observed
from
HCB
exposure,
there
was
an
effect
on
the
reproductive
capability
in
the
Daphnia,
with
a
49
percent
and
41
percent
reduction
in
the
numbers
of
juveniles
produced
and
the
total
biomass
of
juveniles,
respectively,
that
were
produced
during
the
time
span
of
the
experiment.
Niimi
and
Cho
(
1980)
similarly
noted
that
uptake
via
diet
might
be
more
important
than
direct
uptake
for
higher
trophic­
level
organisms,
particularly
in
waters
where
HCB
concentrations
are
low.
Based
on
HCB's
ability
to
bioaccumulate
in
aquatic
organisms,
Fuchsman
et
al.
(
1998)
suggested
that
concentrations
of
HCB
in
surface
water
and
sediment
may
pose
a
risk
to
higher
trophic­
level
organisms
such
as
piscivorous
wildlife
and
recommended
that
HCB
studies
focus
on
higher­
level
species.

Several
studies
have
been
conducted
to
determine
the
equilibration
time
for
HCB
in
aquatic
organisms
relative
to
the
surrounding
surface
water
concentration.
According
to
EPA
(
1988),
HCB
concentrations
in
the
tissues
of
many
aquatic
organisms
appear
to
equilibrate
very
slowly
with
concentrations
in
the
surrounding
water.
Oliver
and
Niimi
(
1983)
reported
that
equilibrium
had
not
yet
occurred
after
119­
day
exposure
of
rainbow
trout
to
HCB
at
a
concentration
of
0.00032

g/
L,
at
which
time
the
bioconcentration
factor
(
BCF)
was
12,000.
HCB
uptake
in
fathead
minnows
was
rapid
during
the
first
week
of
exposure,
then
residues
in
the
minnows
gradually
increased
through
115
days.
HCB
residues
in
the
minnows
represented
a
BCF
of
16,200
at
32
days
exposure
and
a
BCF
of
45,700
at
115
days
exposure
(
Veith
et
al.,
1979).
According
to
EPA
(
1988),
HCB
has
been
shown
to
have
a
relatively
long
half­
life
in
salmonids,
particularly
the
liver.
Rainbow
trout
that
were
fed
HCB
and
maintained
at
a
water
temperature
of
15oC
had
a
half­
life
ranging
from
224
to
770
days
(
EPA,
1988).
The
high
accumulation
levels
of
HCB
in
the
fish
tissue
can
be
explained
by
the
efficient
uptake
of
HCB
from
water
and
food
due
to
its
high
octanol­
water
partition
coefficient
combined
with
the
slow
rate
of
conversion
of
HCB
in
fish
tissue
to
water
soluble
metabolites
and
subsequent
elimination.

The
equilibration
data
provided
in
EPA
(
1988)
are
not
consistent
with
Nebeker
et
al.'
s
(
1989)
investigation.
Nebeler
et
al.
(
1989)
that
found
that
the
equilibration
time
for
HCB
was
somewhat
shorter
in
duration.
Nebeker
et
al.
(
1989)
determined
that
a
period
of
28
days
was
adequate
for
the
equilibration
of
HCB
in
the
water
flea
and
fathead
minnow
tissue
relative
to
ambient
water
concentrations.
They
also
noted
that
the
tissue
concentration
in
the
test
organisms
dropped
rapidly
following
removal
of
the
HCB
from
the
water.
Also,
Koneman
and
Van
Leeuwen
(
1980)
found
an
apparent
steady­
state
between
HCB
residues
in
guppy
tissue
and
surrounding
water
concentration
that
occurred
within
7
days
(
with
a
BCF
of
approximately
16,000).
Differences
in
ambient
water
concentrations
of
HCB
and
other
test
conditions
may
have
contributed
to
the
difference
in
HCB
equilibration
time
observed
for
fathead
minnows
by
Veith
et
al.
(
1979)
of
115
days
compared
with
Nebeker
et
al.'
s
(
1989)
result
of
28
days.
8
If
the
equilibration
time
for
HCB
in
certain
organisms
is
of
long
duration
as
suggested
by
Veith
et
al.
(
1979),
there
is
a
possibility
that
certain
chronic
toxicity
studies
listed
in
Table
1
such
as
Spehar's
(
1986)
90
day
study
of
rainbow
trout
and
Ahmad
et
al
(
1984)
and
Carlson
and
Kosian's
(
1987)
32
day
study
of
the
fathead
minnow
may
not
be
of
sufficient
duration
to
determine
accurate
chronic
threshold
values.
Chronic
toxicity
tests
with
longer
observation
periods
are
needed
to
ensure
that
the
test
organisms
are
in
equilibrium
with
the
concentration
of
HCB
in
the
surrounding
surface
water
and
sediment.

D.
Conclusions
Concerning
Aquatic
Toxicity
of
Hexachlorobenzene
Based
on
reviewed
scientific
literature,
surface
water
concentrations
of
HCB
of
6

g/
L
(
the
solubility
limit
for
this
compound)
do
not
appear
to
result
in
mortality
to
any
tested
species.
Surface
water
concentrations
of
HCB
at
or
below
the
solubility
limit
also
have
a
limited
potential
to
cause
sublethal
toxic
effects
in
aquatic
organisms
for
both
subchronic
and
chronic
exposure.
Although
certain
studies
did
identify
adverse
effects
from
exposure
to
HCB
at
concentrations
less
than
6

g/
L,
including
liver
damage
in
largemouth
bass
(
Laseter
et
al.,
1976)
and
reduced
growth
rates
in
snails
and
crabs
(
Baturo
et
al.,
1995;
and
Mortimer
and
Connell,
1995,
respectively),
the
majority
of
the
data
compiled
in
Appendices
1­
4
do
not
indicate
adverse
effects
to
aquatic
organisms
for
this
range
of
concentrations.
Similarly,
toxicity
studies
did
not
identify
sublethal
adverse
effects
to
benthic
macroinvertebrates
from
exposure
to
HCB
in
sediment.

Some
studies
indicate
that
HCB
concentrations
in
the
tissues
of
aquatic
organisms
appear
to
equilibrate
very
slowly
with
concentrations
in
the
water.
Oliver
and
Niimi
(
1983)
and
Veith
et
al.
(
1979)
found
that
HCB
residue
concentrations
continued
to
increase
in
rainbow
trout
and
fathead
minnows,
respectively,
following
100
days
of
exposure
to
the
compound.
If
HCB
does
require
a
relatively
long
duration
to
reach
equilibrium,
it
is
possible
that
chronic
toxicity
tests
for
certain
species
(
e.
g.,
rainbow
trout
and
fathead
minnow)
may
not
be
of
sufficient
duration
to
accurately
determine
threshold
concentrations.
Also,
because
HCB
tends
to
bioaccumulate
in
aquatic
organisms,
the
potential
for
adverse
effects
to
higher
trophic
level
species,
such
as
piscivorous
wildlife,
is
a
concern.

IV.
TOXICITY
OF
HEXACHLOROBENZENE
TO
TERRESTRIAL
ANIMALS
Adverse
effects
to
individual
wildlife
species
from
HCB
have
been
documented
in
laboratory
studies.
Using
the
results
of
these
studies
to
estimate
effects
on
wild
populations
has
limitations
because
the
exposure
route,
medium
of
administration,
and
exposure
duration
of
HCB
for
wild
animals
usually
will
differ
from
that
of
laboratory
animals.
Using
these
studies
to
assess
effects
on
wild
species
assumes
that
the
wild
species
are
comparably
sensitive
to
HCB
as
laboratory
species;
however,
some
species
may
be
significantly
more
or
less
sensitive
than
laboratory
species.
The
methodologies
for
predicting
the
effects
of
chemicals
on
terrestrial
wildlife
populations
and
ecosystems
are
still
in
development.
Thus,
in
the
absence
of
actual
predictive
methods
for
the
population
level,
measures
of
the
effects
of
HCB
on
reproduction
9
are
currently
the
most
useful
indicators
of
possible
effects
on
the
populations
of
the
species
in
the
wild.

Toxicity
data
for
terrestrial
animals
were
compiled
primarily
from
EPA's
Ecotox
database
(
EPA,
2001),
EPA's
Health
Assessment
Document
for
Chlorinated
Benzenes
(
EPA,
1985),
the
Agency
of
Toxic
Substances
and
Disease
Registry
Toxicological
Profile
for
hexachlorobenzene
(
ATSDR,
2000)
and
miscellaneous
sources
such
as
journal
articles
and
technical
reports.
The
toxicity
data
from
these
four
general
sources
have
been
compiled
in
Appendices
5,6,7,
and
8,
respectively.

A.
Toxicity
Assessment
for
Invertebrates
Toxicity
data
for
terrestrial
invertebrates
were
somewhat
limited.
Neuhauser
(
1985)
conducted
an
acute
toxicity
test
involving
exposure
of
earthworms
(
Eisenia
fetida)
to
HCB
applied
in
a
water
carrier
(
single
dose)
with
a
test
duration
of
48
hours.
As
shown
in
Appendix
5,
test
results
indicated
that
the
LC50
exceeded
1000

g/
cm2.
Similarly,
Ballhorn
(
1984)
found
that
earthworms
exposed
to
HCB
in
artificial
soil
for
a
duration
of
28
days
had
an
LC50
greater
than
1000
mg/
kg
(
Appendix
5).

B.
Toxicity
Assessment
for
Birds
and
Bird
Eggs
The
acute
toxicity
of
HCB
in
birds
was
evaluated
in
several
studies
and
toxicity
data
for
certain
endpoints
(
LC50,
NOEC)
are
provided
in
Table
2
below.

Table
3.
Selected
Avian
Toxicity
Values
Species
Test
Method/
Duration
Endpoint
 
Concentration
Reference
Acute
Toxicity
Ring­
necked
Pheasant
(
Phasianus
colchicus)
5
Days
LC50
 
617
ppm
Hill
et
al.,
1975
Japanese
Quail
5­
day
LC50
 
568
ppm
Hill,
E.
and
M.
Camardese,
1986
Bobwhite
Quail
NR
LD50
­­
575
mg/
kg
HSDB,
1995
Mallard
Duck
NR
LD50
­­
1450
mg/
kg
HSDB,
1995
Mallard
Duck
(
Anas
platyrhynchos)
8
Days
LC50
>
5000
ppm
Hill
et
al.,
1975
Herring
gulls
(
Larus
argentatus)
Egg
injection
study
LD50
 
4300

g/
kg
(
embryo
mortality)
Boersma
et
al.,
1986
Herring
gulls
(
Larus
argentatus)
Egg
injection
study
1500

g/
kg
 
significant
reduction
in
embryonic
weight
Boersma
et
al.,
1986
Chronic
Toxicity
Japanese
Quail
(
Coturnix
japonica)
90
Days
LOEL
 
20
ppm
(
egg
volume)
Vos,
1971
Japanese
Quail
(
Coturnix
japonica)
90
Days
NOEL
 
5
ppm
(
egg
volume)
Vos,
1971
10
Japanese
Quail
(
Coturnix
japonica)
90
Days
LOEL
 
5
ppm
(
morphology,
organ
weight)
Vos,
1971
Japanese
Quail
(
Coturnix
japonica)
90
Days
NOEL
 
1
ppm
(
morphology,
organ
weight)
Vos,
1971
Hexachlorobenzene
is
considered
slightly
to
moderately
toxic
to
bird
species.
As
shown
in
Table
2,
LD50
values
for
evaluated
avian
receptors
generally
ranged
between
568
mg/
kg
and
1450
mg/
kg.
Hill
et
al.
(
1975)
found
that
mallard
ducks
exposed
to
HCB
for
8
days
did
not
exhibit
any
mortality
at
a
concentration
of
5000
ppm
(
LC50
>
5000
ppm).
This
endpoint
was
somewhat
higher
than
other
reported
values,
possibly
the
result
of
interspecies
variation
or
differences
in
specific
test
conditions.
In
regard
to
acute
non­
lethal
effects,
Carpenter
et
al.,
(
1985a)
reported
hepatic
toxicity
and
porphyria
in
Japanese
quail
treated
orally
with
500
mg/
kg/
day
hexachlorobenzene
for
5
days
(
Appendix
5),
with
most
changes
occurring
after
the
first
dose
of
hexachlorobenzene.
Carpenter
et
al.
(
1985b)
also
conducted
a
sub­
chronic
study
of
the
effects
of
hexachlorobenzene
Japanese
quail.
They
determined
that
quail
exposed
to
HCB
for
17
days
at
a
dose
of
100
mg/
kg
per
day
exhibited
alterations
in
porphyrin
level
and
changes
in
organ
weight
related
to
total
body
weight
(
Appendix
5).

Chronic
toxicity
data
for
avian
receptors
in
the
reviewed
literature
were
limited
to
Vos
et
al.'
s
(
1971)
investigation
of
Japanese
quail
(
Coturnix
japonica)
exposed
to
HCB.
The
investigation
involved
dietary
administration
of
0,
1,
5,
20
or
80
mg/
kg
HCB
in
the
diet
for
90
days
(
Vos.
et
al.,
1971).
They
determined
the
NOEL
to
be
1
mg/
kg,
at
higher
concentrations
liver
damage,
porphyrin
excretion
and
alteration
of
enzyme
levels
increased
in
a
dose­
related
fashion.
Vos
et
al.,
(
1971)
also
noted
that
there
was
also
a
dose­
related
decrease
in
the
hatchability
of
eggs,
especially
in
groups
treated
at
20
mg/
kg
for
90
days.

Notably,
a
significant
portion
of
HCB
residue
in
bird
tissue
is
typically
excreted
in
egg
yolks
following
exposure
to
the
compound
(
Extension
Toxicology
Network,
2001).
As
a
result,
studies
on
the
effects
of
HCB
on
wild
birds
have
often
focused
on
the
accumulation
of
the
compound
in
eggs
and
its
subsequent
effects
on
embryo
survival
and
reproductive
parameters.
Egg
hatchability
and
embryo
survival
appear
to
be
two
of
the
more
sensitive
endpoints
for
birds
exposed
to
HCB.
Boersma
et
al.,
(
1986)
conducted
an
egg
injection
study
and
found
a
significant
increase
in
embryo
mortality
with
increasing
dose
of
HCB
in
herring
gulls
(
Larus
argentatus),
and
calculated
the
LD50
to
be
4300

g/
kg
(
4.3
mg/
kg)
as
shown
in
Table
2.
Boersma
et
al.
(
1986)
also
noted
that
an
egg
concentration
of
1500
µ
g/
kg
(
1.5
mg/
kg)
resulted
in
a
significant
reduction
in
embryonic
weight
(
Table
2).

Weseloh
et
al.
(
1983)
also
reported
a
link
between
HCB
egg
concentration
and
successful
avian
reproduction.
They
determined
that
95
percent
of
Double­
breasted
Comorants'
eggs
(
in
a
total
sample
size
of
18
eggs)
with
a
mean
HCB
concentration
of
0.01
ppm
wet
weight
failed
to
hatch,
resulting
in
reproductive
failure.
Jarman
et
al.
(
1996)
evaluated
the
relationship
between
organochlorine
levels
in
prairie
falcon
eggs
and
egg
hatchability
for
falcon
eggs
collected
between
1989
and
1991.
Their
study
results
indicate
that
the
high
HCB
levels
observed
in
the
falcon
eggs
combined
with
the
possible
additive
effect
of
DDE
may
have
had
a
deleterious
effect
on
the
hatching
success
of
prairie
falcons
in
portions
of
the
study
area.
11
C.
Toxicity
Assessment
for
Mammals
The
acute
toxicity
of
HCB
in
mammals
was
evaluated
in
several
studies
and
toxicity
data
for
certain
endpoints
(
LC50,
NOEC)
are
provided
in
Table
3
below.

Table
4.
Selected
Mammalian
Toxicity
Values
Species
Test
Method/
Duration
Endpoint
 
Concentration
Reference
Acute/
Subchronic
Toxicity
Deer
Mouse
(
Peromyscus
maniculatus)
3
Days
(
dosed
one
time
per
study
period)
LD50
 
710
mg/
kg
Schafer
and
Bowles,
1985
Rat
not
reported
LD50
 
3,500
mg/
kg
Savitskii,
1964;
1965
Mouse
not
reported
LD50
 
4,000
mg/
kg
Savitskii,
1964;
1965
Rabbit
not
reported
LD50
 
2,600
mg/
kg
Savitskii,
1964;
1965
Cat
not
reported
LD50
 
1,700
mg/
kg
Savitskii,
1964;
1965
Rat
Fo
to
F1a
and
F1b
generations
21­
day
LD­
50
values
for
pups
were
100
and
140
mg/
kg
for
the
F1a
and
F1b
generations,
respectively
Kitchin
et
al.,
1982
Rat
13
Weeks
19
mg/
kg/
day
 
lethality
(
4
of
9
individuals
died)
Den
Besten
et
al.,
1993
Rat
5
Weeks,
(
dosed
1x
per
day,
5
days
per
week)
LOAEL
 
1
mg/
kg/
day
(
increased
liver
weight)
Andrews
et
al.,
1988
Rat
80
days
LOAEL
 
100
mg/
kg/
day
(
increased
hepatic
porphyrins)
Cantoni
et
al.,
1990
Rat
8
Weeks
(
1x
per
day)
LOAEL
 
1000
mg/
kg/
day
(
increased
liver
weight
and
hepatic
porphyrins)
Kleiman
de
Pisarev
et
al.,
1990
Rat
Chronic/
sub­
chronic
 
30
days
LOAEL
 
1
g/
kg
BW­
day
(
1000
mg/
kg
BW­
day)
 
hepatotoxicity,
porphyria,
alteration
of
oestrus
cycles.
Alvarez,
2000
Chronic
Toxicity
Rat
2
Generations
NOAEL
 
2
mg/
kg/
day
(
reproductive
effects)
Arnold
et
al.,
1985
Rat
Chronic
(>
247
days)
NOAEL
 
1,600

g/
kg
BW­
day
(
reproductive
effects)
Grant
et
al.,
1977
Rat
2
Generations
LOAEL
 
0.016
mg/
kg/
day
(
developmental
effects,
lymphocytosis
and
fibrosis
of
the
liver)
Arnold
et
al.,
1985
Mink
Chronic
(
47
weeks)
LOAEL
 
1
ppm
(
0.16
mg/
kg
BW­
day)
 
reduced
birth
weight
and
increased
mortality
Bleavins
et
al.,
1984
12
The
acute
lethality
of
ingested
HCB
in
terrestrial
mammals
is
relatively
low.
As
shown
in
Table
3,
Savitskii
(
1964;
1965)
determined
that
oral
LD50
values
for
mammals
exposed
to
HCB
ranged
from
1,700
mg/
kg
for
cats
to
4,000
mg/
kg
for
mice;
the
duration
for
these
studies
was
not
reported.
In
a
more
recent
study,
Schafer
and
Bowles
(
1985)
determined
that
the
LD50
for
deer
mice
exposed
to
HCB
for
3
days
was
710
mg/
kg.

Lethal
levels
in
mammalian
studies
are
progressively
lower
as
the
exposure
duration
is
increased.
For
example,
as
shown
in
Table
3,
Kitchin
et
al.
(
1982)
found
that
rats
exposed
to
HCB
in
utero
and
for
21
days
following
birth
had
an
LD50
of
100
mg/
kg
per
day
and
Den
Besten
et
al.
(
1993)
found
that
4
of
9
Wistar
rats
died
following
exposure
to
19
mg/
kg
hexachlorobenzene
per
day
for
a
duration
of
thirteen
weeks.
Study
data
indicate
that
female
animals
are
more
susceptible
to
HCB­
induced
mortality
than
males
(
ATSDR,
2000).

Chronic
toxicity
effects
of
HCB
in
mammals
have
been
provided
in
Appendices
6,7
and
8
and
are
summarized
in
Table
3.
Some
of
the
more
conservative
endpoints
are
associated
with
reproductive
and
developmental
effects
as
shown
in
Table
3.
Arnold
et
al.
(
1985)
determined
that
2
generations
of
rats
exposed
to
HCB
had
a
NOAEL
of
2
mg/
kg/
day
for
reproductive
effects.
Similarly,
Grant
et
al.
(
1977)
found
that
rats
exposed
to
HCB
for
a
duration
grater
than
247
days
had
a
NOAEL
of
1.6
mg/
kg/
day
for
reproductive
effects.
For
developmental
effects,
Arnold
et
al.
(
1985)
found
that
rats
exposed
to
HCB
for
2
generations
had
a
LOAEL
of
0.016
mg/
kg/
day
for
lymphocytosis
and
fibrosis
of
the
liver
and
Bleavins
et
al.
(
1984)
determined
that
mink
exposed
to
HCB
for
47
weeks
had
a
LOAEL
of
0.16
mg/
kg/
day
for
reduced
pup
birth
weight
and
increased
mortality.

D.
Conclusions
Concerning
Terrestrial
Toxicity
of
Hexachlorobenzene
HCB
is
considered
slightly
to
moderately
toxic
to
bird
species.
The
lowest
LC50
dose
for
avian
receptors
was
575
mg/
kg
for
bobwhite
quail.
Sublethal
effects
reported
for
HCB
include
hepatic
toxicity,
porphyria
and
alteration
of
enzyme
levels.
The
lowest
NOAEL
for
HCB
exposure
to
birds
is
1
ppm
(
mg/
kg)
for
Japanese
quail
over
a
90­
day
period.

Scientific
literature
indicates
that
HCB
absorbed
in
avian
tissue
tends
to
be
excreted
in
bird
eggs.
As
a
result,
several
laboratory
and
field
investigations
have
focused
on
this
toxicological
endpoint.
The
lowest
reported
endpoint
for
HCB
in
bird
eggs
in
laboratory
studies
was
1.5
mg/
kg
in
herring
gull
eggs,
which
caused
a
significant
reduction
in
embryonic
weight.
Notably,
HCB
concentrations
observed
in
the
eggs
of
sea
birds
and
raptors
from
a
number
of
locations
from
around
the
world
approach
this
level,
suggesting
that
HCB
could
potentially
cause
an
adverse
affect
on
the
reproduction
of
bird
species.

The
acute
lethality
of
ingested
HCB
in
terrestrial
mammals
is
relatively
low.
Reviewed
LD50
values
range
from
710
mg/
kg
for
deer
mice
(
over
a
3­
day
exposure
period)
to
4,000
mg/
kg
for
mice
(
species
and
exposure
duration
not
reported).
The
toxicity
of
HCB
in
mammal
studies
increases
with
increasing
exposure
duration.
Sublethal
effects
to
mammals
from
HCB
exposure
13
include
liver
damage,
alteration
of
oestrus
cycles
and
other
reproductive
effects
and
reduced
birth
weight.

For
chronic
exposure,
the
lowest
reported
NOAEL
was
1.6
mg/
kg­
BW
per
day
in
rats
(
over
a
duration
greater
than
247
days)
for
reproductive
effects.
Another
sensitive
endpoint
is
the
reduction
of
birth
weight
and
increased
mortality
in
mink
offspring
following
chronic
exposure
to
1
ppm
HCB
(
0.16
mg/
kg
BW­
day)
for
a
duration
of
47
weeks.
Field
observations
of
HCB
concentrations
in
fish
tissue
at
a
number
of
sites
worldwide
that
are
within
an
order
of
magnitude
of
1
ppm,
suggesting
that
HCB
has
the
potential
to
cause
adverse
effects
in
mink
and
perhaps
other
fish­
eating
mammals.

V.
Ecological
Risk
Assessment
Risk
assessment
integrates
the
results
of
the
exposure
and
ecotoxicity
data
to
evaluate
the
likelihood
of
adverse
ecological
effects.
One
method
of
integrating
the
results
of
exposure
and
ecotoxicity
data
is
called
the
quotient
method.
For
this
method,
risk
quotients
(
RQs)
are
calculated
by
dividing
exposure
estimates
by
ecotoxicity
values,
both
acute
and
chronic:

RQ
=
EXPOSURE/
TOXICITY
RQs
are
then
compared
to
AD's
levels
of
concern
(
LOCs).
These
LOCs
are
criteria
used
by
OPP
to
indicate
potential
risk
to
nontarget
organisms
and
the
need
to
consider
regulatory
action.
The
criteria
indicate
that
a
pesticide
used
as
directed
has
the
potential
to
cause
adverse
effects
on
nontarget
organisms.
LOCs
currently
address
the
following
risk
presumption
categories:
(
1)
acute
high
­
potential
for
acute
risk
is
high
regulatory
action
may
be
warranted
in
addition
to
restricted
use
classification;
(
2)
acute
restricted
use
­
the
potential
for
acute
risk
is
high,
but
this
may
be
mitigated
through
restricted
use
classification;
(
3)
acute
endangered
species
­
the
potential
for
acute
risk
to
endangered
species
is
high,
and
regulatory
action
may
be
warranted,
and
(
4)
chronic
risk
­
the
potential
for
chronic
risk
is
high,
and
regulatory
action
may
be
warranted.
Currently,
AD
does
not
perform
assessments
for
chronic
risk
to
plants,
acute
or
chronic
risks
to
nontarget
insects,
or
chronic
risk
from
granular/
bait
formulations
to
mammalian
or
avian
species.

The
ecotoxicity
test
values
(
i.
e.,
measurement
endpoints)
used
in
the
acute
and
chronic
risk
quotients
are
derived
from
the
results
of
required
studies.
Examples
of
ecotoxicity
values
derived
from
the
results
of
short­
term
laboratory
studies
that
assess
acute
effects
are:
(
1)
LC50
(
fish
and
birds)
(
2)
LD50
(
birds
and
mammals
(
3)
EC50
(
aquatic
plants
and
aquatic
invertebrates)
and
(
4)
EC25
(
terrestrial
plants).
Examples
of
toxicity
test
effect
levels
derived
from
the
results
of
longterm
laboratory
studies
that
assess
chronic
effects
are:
(
1)
LOEC
(
birds,
fish,
and
aquatic
invertebrates)
(
2)
NOEC
(
birds,
fish
and
aquatic
invertebrates)
and
(
3)
MATC
(
fish
and
aquatic
invertebrates).
For
birds
and
mammals,
the
NOEC
value
is
used
as
the
ecotoxicity
test
value
in
assessing
chronic
effects.
Other
values
may
be
used
when
justified.
Generally,
the
MATC
(
defined
as
the
geometric
mean
of
the
NOEC
and
LOEC)
is
used
as
the
ecotoxicity
test
value
in
14
assessing
chronic
effects
to
fish
and
aquatic
invertebrates.
However,
the
NOEC
is
used
if
the
measurement
endpoint
is
production
of
offspring
or
survival.

Risk
presumptions,
along
with
the
corresponding
RQs
and
LOCs
are
tabulated
below.

Risk
Presumptions
for
Terrestrial
Animals
(
Birds
and
Wild
Mammals)

Risk
Presumption
RQ
LOC
Acute
High
Risk
EEC1/
LC50
or
LD50/
sq
ft2
or
LD50/
day3
0.5
Acute
Restricted
Use
EEC/
LC50
or
LD50/
sq
ft
or
LD50/
day
(
or
LD50
<
50
mg/
kg)
0.2
Acute
Endangered
Species
EEC/
LC50
or
LD50/
sq
ft
or
LD50/
day
0.1
Chronic
Risk
EEC/
NOEC
1
1
abbreviation
for
Estimated
Environmental
Concentration
(
ppm)
on
avian/
mammalian
food
items
2
mg
/
ft
LD50
*
wt.
of
bird
2
3
mg
of
toxicant
consumed
/
day
LD50
*
wt.
of
bird
Risk
Presumptions
for
Aquatic
Animals
Risk
Presumption
RQ
LOC
Acute
High
Risk
EEC1/
LC50
or
EC50
0.5
Acute
Restricted
Use
EEC/
LC50
or
EC50
0.1
Acute
Endangered
Species
EEC/
LC50
or
EC50
0.05
Chronic
Risk
EEC/
MATC
or
NOEC
1
1
EEC
=
(
ppm
or
ppb)
in
water
15
Risk
Presumptions
for
Plants
Risk
Presumption
RQ
LOC
Terrestrial
and
Semi­
Aquatic
Plants
Acute
High
Risk
EEC1/
EC25
1
Acute
Endangered
Species
EEC/
EC05
or
NOEC
1
Aquatic
Plants
Acute
High
Risk
EEC2/
EC50
1
Acute
Endangered
Species
EEC/
EC05
or
NOEC
1
1
EEC
=
lbs
ai/
A
2
EEC
=
(
ppb/
ppm)
in
water
A.
Exposure
and
Risk
to
Nontarget
Terrestrial
Animals
Terrestrial
organisms
may
be
exposed
to
HCB
through
direct
contact
with
treated
lumber
and
contact
with
HCB
in
soil
that
has
leached
from
treated
lumber.
Exposure
and
toxicity
data
for
direct
contact
with
treated
lumber
were
not
readily
available
for
terrestrial
organisms.
As
a
result,
the
terrestrial
assessment
was
based
on
exposure
to
HCB
in
soil.

Risks
to
mammalian
and
avian
species
were
evaluated
using
a
simple
terrestrial
food
web
model.
To
estimate
exposure
for
receptor
species,
it
was
assumed
that
HCB
leaching
occurs
from
treated
wood
into
the
surrounding
soil.
The
leached
HCB
in
soil
was
then
assumed
to
be
taken
up
by
terrestrial
vegetation.
Based
on
these
conditions,
the
model
assumed
that
birds
and
mammals
would
be
exposed
to
HCB
through
ingestion
of
vegetation
and
incidental
ingestion
of
soil.

The
estimated
environmental
concentration
(
EEC)
for
soil
used
in
the
terrestrial
risk
assessment
was
determined
by
modeling
described
in
U.
S.
EPA's
"
Environmental
Fate
Modeling
of
Hexachlorobenzene
in
Technical
Grade
Pentachlorophenol"
(
EPA,
2002).
Using
the
state
of
Ohio
to
represent
a
typical
regional
environmental
setting,
U.
S.
EPA
calculated
the
amount
of
HCB
released
from
an
estimated
1,500,000
utility
poles
within
the
state.
Based
on
the
calculated
release
rates
of
HCB
via
volatilization,
wood
erosion,
and
leaching,
the
fate
and
transport
of
HCB
releases
was
modeled
using
the
Level
III
model.
The
HCB
released
was
assumed
to
partition
into
the
air,
soil,
water,
and
sediments.
According
to
the
results,
terrestrial
organisms
would
be
exposed
to
2.58
mg/
kg
of
HCB
in
the
soil.
16
Table
5.
Summary
of
HCB
Concentrations
in
Air,
Water,
Sediments,
and
Soils,
as
Calculated
in
U.
S.
EPA
(
2002).
Concentration
Air
(
g/
m3)
1.39E­
12
Water
(
mg/
L)
1.76E­
11
Sediment
(
µ
g/
g)
8.13E­
08
Soils
(
mg/
kg)
2.58E­
08
The
methods
used
to
estimate
the
potential
dose
of
HCB
to
avian
and
mammalian
species
through
the
ingestion
of
vegetation
and
soil
are
provided
in
Appendix
9.
The
results
of
the
exposure
assessment
for
avian
and
mammalian
species
have
been
provided
in
Tables
A1
and
A2
(
in
Appendix
9).

Generally,
the
hierarchy
of
sources
for
selecting
which
toxicity
value
to
use
when
more
than
one
was
identified
is:
1)
core
studies
from
the
pesticide
database,
2)
supplemental
studies
from
the
pesticide
database,
3)
ECOTOX
studies,
and
4)
studies
from
the
open
literature.
However,
since
the
original
studies
were
not
available
for
review,
when
more
than
one
toxicity
reference
value
was
available,
the
toxicity
data
that
had
the
most
complete
study
information
(
e.
g.,
reported
study
duration
or
toxicity
endpoint
defined)
and
had
the
lowest
toxicity
value
was
selected
in
order
to
be
conservative.

i.
Birds
As
shown
in
Table
A1,
the
risks
to
avian
species
from
exposure
to
HCB
in
soil
as
a
result
of
HCB
leaching
from
treated
lumber
does
not
appear
to
be
significant.
The
RQ
of
4.3E­
13
for
acute
exposure
to
HCB
was
much
less
than
all
LOCs.
The
chronic
RQ
for
HCB,
2.5E­
10,
was
also
much
less
than
the
LOC.

ii.
Mammals
Based
on
the
results
in
Table
A2,
exposure
to
concentrations
of
HCB
in
soil
would
not
result
in
adverse
effects
to
mammalian
species.
The
RQ
for
acute
exposure
to
HCB,
1.3E­
12,
was
below
all
the
defined
LOCs.
Similarly,
the
chronic
RQ
for
HCB
estimated
using
the
predicted
soil
concentration,
9.9E­
10,
was
much
less
than
any
of
the
LOCs.

iii.
Conclusions
of
the
Risk
Assessment
for
Birds
and
Mammals:

The
results
indicate
that
risks
to
birds
and
mammals
exposed
to
concentrations
of
HCB
in
soil
is
low.
It
should
be
noted,
however,
that
the
risk
assessment
was
only
based
on
exposure
to
HCB
components
in
soil.
A
quantitative
assessment
of
the
risks
to
birds
and
mammals
from
direct
contact
with
HCB­
treated
lumber
was
not
conducted
due
to
the
lack
of
exposure
and
toxicity
data
available.
As
a
result,
the
potential
risks
from
direct
contact
with
HCB­
treated
wood
were
not
evaluated.
Additional
uncertainties
associated
with
the
assessment
are
discussed
in
the
Uncertainty
section
of
the
report.
17
iv.
Insects
Exposure
to
HCB
in
treated
lumber
and
the
surrounding
soil
may
result
in
adverse
effects
to
insects
such
as
honey
bees.
The
potential
risks
to
insects
could
not
be
quantitatively
evaluated,
however,
since
toxicity
and
exposure
data
for
insects
were
not
readily
available.

B.
Exposure
and
Risk
to
Nontarget
Freshwater
and
Marine/
Estuarine
Aquatic
Organisms
Nontarget
freshwater
and
marine/
estuarine
aquatic
organisms
could
potentially
be
exposed
to
HCB
leached
from
treated
wood
into
the
aquatic
environment.
RQs
were
calculated
for
acute
and
chronic
effects
on
aquatic
organisms
using
EECs
for
surface
water.
The
HCB
EEC
in
water,
1.76E­
11
mg/
L
was
determined
by
modeling
the
fate
and
transport
of
HCB
released
from
utility
poles
in
the
environment
and
detailed
in
U.
S.
EPA
(
2002).

Using
this
EEC,
RQ
values
were
calculated
and
the
results
are
presented
in
the
following
tables.
Both
acute
and
chronic
RQs
were
estimated
using
the
predicted
surface
water
EEC.
As
previously
noted,
when
more
than
one
toxicity
reference
value
was
available,
generally,
the
lowest
value
was
selected
in
order
to
be
conservative.

i.
Freshwater
Fish
Acute
and
chronic
RQs
for
freshwater
fish
are
tabulated
below.
The
lowest
toxicity
(
LC50)
values
were
used
for
HCB
in
order
to
be
conservative.

Table
6.
HCB
Acute
and
Chronic
Risk
Quotients
for
Freshwater
Fish
Acute
Exposures
Chronic
Exposures
LC50
(
mg/
L)
EEC
(
mg/
L)
Acute
RQ
(
EEC/
LC50)
NOEC
(
mg/
L)
EEC
(
mg/
L)
Chronic
RQ
(
EEC/
NOEC)
HCB
6.4
(
Mayer
and
Ellerseick,
1986)
1.76E­
11
2.8E­
12
(
a)
3.68
(
Spehar,
1986)
1.76E­
11
4.8E­
12
(
a)

a
=
none
of
the
LOCs
have
been
exceeded.

Results
indicate
that
none
of
the
RQs
exceed
any
LOC.

ii.
Freshwater
Invertebrates
Acute
and
chronic
RQs
for
aquatic
invertebrates
have
been
tabulated
below.
The
lowest
toxicity
(
LC50)
value
was
used
in
order
to
be
conservative.
18
Table
7.
HCB
Acute
and
Chronic
Risk
Quotients
for
Freshwater
Invertebrates
Acute
Exposures
Chronic
Exposures
EC50
(
mg/
L)
EEC
(
mg/
L)
Acute
RQ
(
EEC/
EC50)
NOEC
(
mg/
L)
EEC
(
mg/
L)
Chronic
RQ
(
EEC/
NOEC)
HCB
0.005
(
Laseter
et
al.,
1976)
1.76E­
11
3.5E­
09
(
a)
0.007
(
Spehar,
1986)
1.76E­
11
2.5E­
09
(
a)

a
=
none
of
the
LOCs
have
been
exceeded.

The
RQs
based
on
the
EEC
values
from
the
leachate
study
indicate
that
the
none
of
the
LOCs
are
exceeded
for
acute
or
chronic
exposures.

iii.
Estuarine
and
Marine
Fish
Acute
risk
quotients
(
RQs)
for
estuarine
and
marine
fish
are
tabulated
below.
The
lowest
toxicity
(
LC50)
value
was
used
in
order
to
be
conservative.

Table
8.
HCB
Acute
Risk
Quotient
for
Estuarine/
Marine
Fish
Acute
Exposures
LC50
(
mg/
L)
EEC
(
mg/
L)
Acute
RQ
(
EEC/
LC50)

HCB
0.14
(
Parrish
et
al.,
1976)
1.76E­
11
1.3E­
10
(
a)
a
=
none
of
the
LOCs
have
been
exceeded.

The
calculation
indicates
that
the
acute
RQ
did
not
exceed
any
LOCs.

No
chronic
toxicity
data
for
marine/
estuarine
fish
or
invertebrate
species
were
available.
Therefore,
risk
quotients
could
not
be
calculated
for
these
organisms.
Acute
fish
toxicity
data
indicate
that
marine/
estuarine
fish
species
are
more
sensitive
to
HCB
than
freshwater
species;
therefore,
the
chronic
risk
posed
to
these
species
by
HCB
may
be
greater
than
the
chronic
risks
posed
to
freshwater
species.
However,
the
chronic
RQs
developed
for
freshwater
species
are
well
below
the
LOCs,
and
even
estimating
a
two
order
of
magnitude
difference
(
based
on
the
freshwater
fish
acute
RQ
compared
to
the
marine/
esutarine
fish
acute
RQ),
the
chronic
RQ
for
marine/
estuarine
species
would
likely
remain
below
the
LOC.

iv.
Conclusions
of
the
Risk
Assessment
for
Aquatic
Animals:

Results
of
the
risk
assessment
indicate
that
the
potential
risk
to
aquatic
animals
exposed
to
HCB
leachate
in
surface
water
would
be
low.
Both
the
acute
and
chronic
RQs
were
significantly
below
their
respective
LOQs
for
each
group
of
aquatic
animals
evaluated.
However,
as
discussed
in
the
hazard
section
of
this
assessment,
HCB
has
a
strong
tendency
to
bioaccumulate,
and
could
therefore
potentially
cause
adverse
effects
to
aquatic
organisms
over
time,
and
could
possibly
cause
adverse
effects
to
higher
trophic
level
organisms
which
consume
HCB­
contaminated
aquatic
organisms.
19
C.
Exposure
and
Risk
to
Plants
i.
Terrestrial
Plants
The
use
of
HCB
as
a
wood
preservative
may
result
in
the
leaching
of
HCB
into
the
surrounding
soil.
However,
no
terrestrial
plant
studies
were
identified
and,
thus,
the
risks
to
terrestrial
plants
could
not
be
quantitatively
evaluated.

ii.
Aquatic
Plants
Acute
risk
quotients
(
RQs)
for
aquatic
plants
are
tabulated
below.
The
lowest
toxicity
value
was
used
in
order
to
be
conservative.

Table
9
HCB
Acute
Risk
Quotient
for
Aquatic
Plants
Acute
Exposures
EC50
(
mg/
L)
EEC
(
mg/
L)
Acute
RQ
(
EEC/
EC50)

HCB
0.003
(
Wong
et
al.,
1984)
1.76E­
11
8.8E­
09
(
a)
a
=
none
of
the
LOCs
have
been
exceeded.

The
RQ
based
on
the
EEC
value
for
HCB
leached
from
utility
poles
is
below
all
LOCs.

VI.
Endangered
Species
The
results
of
the
risk
assessment
indicate
that
threatened
and
endangered
species
would
not
be
expected
to
be
adversely
affected
directly
by
exposure
HCB.
However,
as
discussed
above,
the
strong
tendency
of
HCB
to
bioaccumulate
could
lead
to
secondary
adverse
effects
to
higher
trophic
level
organisms,
or
direct
effects
to
organisms
exposed
to
HCB
over
longer
periods
of
time.

The
Agency
has
developed
a
program
(
the
Endangered
Species
Protection
Program)
to
identify
pesticides
whose
use
may
result
in
adverse
impacts
to
endangered
and
threatened
species,
and
to
implement
mitigation
measures
that
will
eliminate
the
adverse
impacts.
At
present,
the
program
is
being
implemented
on
an
interim
basis
as
described
in
a
Federal
Register
notice
(
54
FR
27984­
28008,
July
3,
1989),
and
is
providing
information
to
pesticide
users
to
help
them
protect
these
species
on
a
voluntary
basis.
As
currently
planned,
the
final
program
will
call
for
label
modifications
referring
to
required
limitations
on
pesticide
uses,
typically
as
depicted
in
countyspecific
bulletins
or
by
other
site­
specific
mechanisms
as
specified
by
state
partners.
A
final
program,
which
may
be
altered
from
the
interim
program,
will
be
described
in
a
future
Federal
Register
notice.
The
Agency
is
not
imposing
label
modifications
at
this
time
through
the
RED.
Rather,
any
requirements
for
product
use
modifications
will
occur
in
the
future
under
the
Endangered
Species
Protection
Program.
20
VII.
Uncertainties
in
the
Risk
Assessment
All
of
the
toxicity
values
used
in
the
assessment
were
from
the
ECOTOX
database
or
the
open
literature
rather
than
submitted
studies
that
meet
the
pesticide
guideline
requirements,
as
HCB
is
not
a
registered
pesticide.
Testing
conditions
in
these
studies
were
likely
quite
variable,
which
causes
greater
variability
across
test
results.

Another
source
of
uncertainty
in
the
toxicity
data
used
in
the
assessment
is
that
most
of
the
toxicity
studies
were
conducted
in
a
laboratory
environment,
and
therefore
may
not
accurately
reflect
field
conditions
experienced
by
wild
organisms.
Toxicity
studies
also
frequently
use
different
terrestrial
or
aquatic
species
than
those
expected
to
occur
in
the
natural
environment.
The
comparative
sensitivity
between
wild
organisms
and
laboratory
organisms
is
not
well­
known,
and,
therefore,
some
wild
species
may
be
substantially
more
sensitive
to
HCB
than
is
indicated
by
the
toxicity
endpoints
used
in
this
assessment.

There
are
several
uncertainties
associated
with
the
exposure
assessment
for
HCB
in
terrestrial
and
aquatic
environments.
The
terrestrial
risk
assessment
assumed
that
birds
and
mammals
were
only
exposed
to
HCB
in
soil,
since
toxicity
and
exposure
data
were
not
readily
available
for
direct
contact
with
pentachlorophenol­
treated
wood.
Likewise,
the
aquatic
risk
assessment
assumed
that
aquatic
organism
exposure
to
HCB
would
occur
only
from
leaching
from
pentachlorophenoltreated
wood.
As
discussed
in
the
separate
Environmental
Fate
Modeling
of
HCB
chapter,
volatilization,
wood
erosion
and
leaching
all
contribute
HCB
to
the
environment.
The
combined
amount
of
HCB
released
from
pentachlorophenol­
treated
wood,
rather
than
the
amount
released
from
one
specific
pathway,
would
provide
a
more
accurate
estimate
of
the
levels
to
which
terrestrial
and
aquatic
organisms
are
exposed.
The
model
also
included
many
assumptions
regarding
the
indicator
species,
such
as
diet,
food
ingestion
rate,
and
home
range
size,
that
may
not
accurately
reflect
actual
conditions
in
the
environment.
21
VIII.
REFERENCES:

Abernethy
S.,
Bobra
A.,
Shiu
W.,
Wells
P.
and
D.
MacKay.
1986.
"
Acute
Lethal
Toxicity
of
Hydrocarbons
and
Chlorinated
Hydrocarbons
to
Two
Planktonic
Crustaceans:
The
Key
Role
of
Organism­
Water
Partitioning."
Aquatic
Toxicology,
8(
3):
163­
174.

Agency
for
Toxic
Substances
and
Disease
Registry
(
ATSDR).
2000.
Toxicological
Profile
for
Hexachlorobenzene.

Ahmad
N.,
Benoit
D.,
Brooke
L.,
Call
D.,
Carlson
A.,
Defoe
D.,
Huot
J.,
Moriarity
A.,
Richter
J.,
Shubat
P.,
Veith
G.
and
C.
Walbridge.
1984.
Aquatic
Toxicity
Tests
to
Characterize
the
Hazard
of
Volatile
Organic
Chemicals
in
Water:
A
Toxicity
Data
Summary
 
Parts
I
and
II.
EPA­
600/
3­
84­
009.
National
Technical
Information
Service,
Springfield,
Virginia.

Alvarez
L.,
Randi
A.,
Alvarez
P.,
Piroli
G.,
Chamson­
Reig
A.,
Lux­
Lantos
V.,
and
D.
Kleiman
de
Pisarev.
2000.
"
Reproductive
effects
of
Hexachlorobenzene
in
Female
Rats,"
Journal
of
Applied
Toxicology,
20:
81­
87.

Andrews
J.,
Courtney
K.
and
W.
Donaldson.
1988.
"
Impairment
of
Calcium
Homeostasis
by
Hexachlorobenzene
(
HCB)
Exposure
in
Fischer
344
Rats."
Journal
of
Toxicology
and
Environmental
Health,
23:
311­
320.

Arnold
D.,
Moodie
C.,
Charbonneau
S.
et
al.
1985.
"
Long­
term
Toxicity
of
Hexachlorobenzene
in
the
Rat
and
the
Effect
of
Dietary
Vitamin
A.
Food
Chemistry
and
Toxicology,
23:
779­
793.

Bailey,
J.,
Krianf,
V.,
Mueller,
and
W.
Hobson.
1980.
Transfer
of
Hexachlorobenzene
and
Polychlorinated
Biphenyls
to
Nursing
Infant
Rhesus
Monkeys:
Enhanced
Toxicity.
Environmental
Research,
21(
1):
190­
196.

Ballhorn,
L.,
Freitag,
D.,
Geyer,
H.,
Quast
I.,
et
al.
1984.
Uberprufung
der
Durchfuhrbarkeit
von
Prufungsvorschriften
und
der
Aussagekraft
der
Stufe
I
and
II
des
E.
Chem
G.
Forschungsbericht
No.
106
04
011/
02,
Umweltbundesamt
Berlin.

Barber
T.,
Fuchsman
P.,
Chappie
D.,
Sferra
J.,
Newton
F.,
and
P.
Sheehan.
1997.
"
Toxicity
of
Hexachlorobenzene
to
Hyalella
azteca
and
Chironomus
tentans
in
Spiked
Sediment
Bioassays."
Environmental
Toxicology
and
Chemistry,
16(
8):
1716­
1720.

Baturo
W.,
Lagadic
L.
and
T.
Caquet.
1995.
"
Growth,
Fecundity
and
Glycogen
Utilization
in
Lymnaea
palustris
Exposed
to
Atrazine
and
Hexachlorobenzene
in
Freshwater
Mesocosms."
Environmental
Toxicology
and
Chemistry,
14(
3):
503­
511.

Beyer,
W.;
Conner,
E.
and
S.
Gerould.
1994.
Estimates
of
Soil
Ingestion
by
Wildlife.
Journal
of
Wildlife
Management,
58:
375­
382.
22
Biggs,
D.
C.,
Rowland,
R.
G.,
and
C.
F.
Wurster.
1979.
Effects
of
Trichloroethylene,
Hexachlorobenzene
and
Polychlorinated
Biphenyls
on
the
Growth
and
Cell
Size
of
Marine
Phytoplankton.
Bulletin
of
Environmental
Contamination
Toxicology,
21(
1­
2):
196­
201.

Bleavins,
M.,
Aulerich,
R.
and
R.
Ringer.
1984.
"
Effect
of
Chronic
Dietary
Hexachlorobenzene
Exposure
on
the
Reproductive
Performance
and
Survivability
of
Mink
and
European
Ferrets,"
Archives
of
Environmental
Contaminant
Toxicology,
13:
357­
365.

Boersma,
D.,
Ellenton,
J.,
and
A.
Yagminas.
1986.
"
Investigation
of
the
Hepatic
Mixed­
Function
Oxidase
System
in
Herring
Gull
Embryos
in
Relation
to
Environmental
Contaminants.
Environmental
Toxicology
and
Chemistry,
5:
309­
318.

Calamari
D.,
Galassi
S.,
Setti
F.
and
M.
Vighi.
1983.
"
Toxicity
of
Selected
Chlorobenzenes
to
Aquatic
Organisms.
Chemosphere,
12:
253­
262.

Call
D.,
Brooke
L.,
Ahmad
N.
and
J.
Richter.
1983.
Toxicity
and
Metabolism
Studies
with
EPA
Priority
Pollutants
and
Related
Chemicals
in
Freshwater
Organisms.
EPA­
600/
3­
83­
095.
National
Technical
Information
Service,
Springfield
VA.

Cantoni,
L.,
Budillon,
G.,
and
R.
Cuomo
et
al.
1990.
Protective
Effect
of
S­
adenosyl­
Lmethionine
in
Hepatic
Uroporphyria,
Evaluation
in
an
Experimental
Model.
Stand
J
Gastro,
25:
1034­
1040.

Carlson
A.,
and
P.
Kosian.
1987.
"
Toxicity
of
Chlorinated
Benzenes
to
Fathead
Minnows
(
Pimephales
promelas).
Archives
Environmental
Contamination
and
Toxicology,
16:
129­
135.

Carpenter,
H.,
Williams,
D.
and
D.
Buhler.
1985a.
"
A
Comparison
of
the
Effects
of
Hexachlorobenzene,
Beta­
Naphthoflavone,
and
Phenobarbital
on
Cytochrome
P­
450
and
Mixed­
Function
Oxidases
in
Japanese
Quail,"
Journal
of
Toxicology
and
Environmental
Health,
15:
93­
108.

Carpenter,
H.,
Harvey,
M.
and
D.
Buhler.
1985b.
"
The
Effect
of
Tetrachlorohydroquinone
on
Hexachlorobenzene­
induced
Porphyria
in
Japanese
Quail,"
Journal
of
Toxicology
and
Environmental
Health,
15:
81­
92.

Caspers
N.,
Hartmann
P.,
Kanne
R.
and
G.
Knoop.
1993.
Bund/
Lander­
Arbeitskreis
"
Qualitatsziele
(
BLAK­
QZ)"
 
Problematik
des
Konzepts
bei
der
Festlegung
von
Zielvorgaben
fur
Oberflachengewasser
am
Beispiel
von
Trichlorethen
und
Hexachlorbenzol.
Umweltwissenschaften
und
Schadstoff­
Forschung,
5(
5):
265­
270.

Cuomo
R.,
Rodino
S.,
and
R.
Rizzoli
et
al.
1991.
Bile
and
Biliary
Lipid
Secretion
in
Rats
with
Hexachlorobenzene­
induced
Porphyria.
Effect
of
S­
adenosyl­
L­
methionine
Administration.
Journal
of
Hepatology,
12:
87­
93.
23
Den
Bensten,
C.,
Bennik
M.,
Bruggeman
I.,
et
al.
1993.
The
Role
of
Oxidative
Metabolism
in
Hexachlorobenzene­
induced
Porphyria
and
Thyroid
Hormone
Homeostasis:
A
Comparison
with
Pentachlorobenzene
in
a
13­
week
Feeding
Study.
Toxicology
and
Applied
Pharmacology,
119:
181­
194.

Dive
D.,
LeClerc
H.,
and
G.
Personne.
1980.
"
Pesticide
Toxicity
on
the
Ciliate
Protozoan
Colpidium
campylum:
Possible
Consequences
of
the
Effect
of
Pesticides
in
the
Aquatic
Environment."
Ecotoxicology
Environmental
Safety,
4:
129­
133.

Extension
Toxicology
Network.
2001.
Cornell
University,
http://
pmep.
cce.
cornell.
edu/
profiles/
extoxnet/

Figueroa
I.
And
M.
Simmons.
1991.
"
Structure­
Activity
Relationships
of
Chlorobenzenes
Using
DNA
Measurement
as
a
Toxicity
Parameter
in
Algae."
Environmental
Toxicology
Chemistry,
10(
3):
323­
329.

Furay
V.
and
S.
Smith.
1995.
"
Toxicity
of
QSAR
of
Chlorobenzenes
in
Two
Species
of
Benthic
Flatfish,
Flounder
(
Platichthys
flesus
L.)
and
Sole
(
Solea
solea
L.)."
Bulletin
of
Environmental
Contamination
and
Toxicology,
54:
36­
42.

Fuchsman
P.,
Barber
T.,
and
P.
Sheehan.
1998.
"
Sediment
Toxicity
Evaluation
for
Hexachlorobenzene:
Spiked
Sediment
Tests
with
Leptocheirus
plumulosus,
Hyalella
azteca,
and
Chironomus
tentans."
Archives
of
Environmental
Contamination
and
Toxicology,
35:
573­
579.

Geyer
H.,
Scheunert
I.
And
F.
Korte.
1985.
"
The
Effects
of
Organic
Environmental
Chemicals
on
the
Growth
of
the
Alga
Scenedesmus
subspicatus:
A
Contribution
to
Environmental
Biology."
Chemosphere,
14:
1355­
1369.

Grant,
D.,
Phillips,
W.
and
Hatina,
G.
1977.
"
Effect
of
Hexachlorobenzene
on
Reproduction
in
the
Rat,"
Archives
of
Environmental
Contamination
and
Toxicology,
5:
207­
216.

Hill,
E.,
Heath,
J.,
Spann,
J.,
and
J.
Williams.
1975.
"
Lethal
Dietary
Toxicities
of
Environmental
Pollutants
to
Birds,"
U.
S.
Fish
and
Wildlife
Service,
Special
Scientific
Report
 
Wildlife,
191:
1­
61.

Hill,
E.
and
M.
Camardese.
1986.
"
Lethal
Dietary
Toxicities
of
Environmental
Contaminants
and
Pesticides
to
Coturnix,"
Technical
Report
Number
2.
U.
S.
Department
of
Interior,
Fish
and
Wildlife
Service,
Washington,
DC,
pp.
6­
55.

Jarman,
W.,
Burns
S.,
Bacon,
C.,
Rechtin,
J.,
DeBenedetti
S.,
Linthicum,
J.
and
B.
Walton.
1996.
"
High
Levels
of
HCB
and
DDE
Associated
with
Reproductive
Failure
in
Prairie
Falcons
(
Falco
mexicanus)
from
California,"
Bulletin
of
Environmental
Contamination
and
Toxicology,
57:
8­
15.
24
Johnson
W.
and
M.
Finley.
1980.
Handbook
of
Acute
Toxicity
of
Chemicals
to
Fish
and
Aquatic
Invertebrates.
Research
Publication
137.
U.
S.
Fish
and
Wildlife
Service,
Washington
DC.,
pp.
44­
45.

Kennedy
S.
and
D.
Wigtield.
1990.
"
Dose­
Response
Relationships
in
Hexachlorobenzeneinduced
Porphyria."
Biochemistry
Pharmacology,
40:
1381­
1388.

Kitchin,
K.
T.,
Linder,
R.
E.,
Scotti,
T.
M.,
Walsh,
I.,
Curley,
A.
D.,
and
D.
Svensguard.
1982.
Offspring
Mortality
and
Maternal
Lung
Pathology
in
Female
Rats
Fed
Hexachlorobenze.
Toxicology,
23(
1):
33­
39.

Kleiman
de
Pisarev
D.,
Sancovich,
H.
and
A.
Ferramola
de
Sancovich.
1995.
"
Hepatic
Indices
of
Thyroid
Status
in
Rats
Treated
with
Hexachlorobenzene."
J
Endocrinology
Investigations,
18:
271­
276.

Kleiman
de
Pisarev
D.,
Rios
de
Molina,
M.
and
L.
San
Martin
de
Viale.
1990.
"
Thyroid
Function
and
Thyroxine
Metabolism
in
Hexachlorobenzene­
induced
Porphyria."
Biochemistry
Pharmacology,
39:
817­
825.

Kleiman
de
Pisarev
D.,
Sancovich,
H.
and
A.
Ferramola
de
Sancovich.
1989.
"
Enhanced
Thyroxine
Metabolism
in
Hexachlorobenzene­
Intoxicated
Rats."
J
Endocrinology
Investigation,
12:
767­
772.

Knie
J.,
Halke
A.,
Juhnke
I.
And
W.
Schiller.
1983.
Results
of
Studies
on
Chemical
Substances
with
Four
Biotests.
Dtsch.
Gewaesserkd.
Mitt,
27(
3):
77­
79.

Kolok
A.,
Groetsch
K.,
and
J.
Oris.
1996.
"
The
Role
of
Water
Ventilation
and
Sediment
Ingestion
on
the
Uptake
of
Hexachlorobenzene
by
Gizzard
Shad
(
Dorosoma
cepedianum)."
Environmental
Toxicology
and
Chemistry,
15(
10):
1760­
1762.

H.
Konemann.
1979.
"
Quantitative
Structure­
Activity
Relationships
for
Kinetics
and
Toxicity
of
Aquatic
Pollutants
and
their
Mixtures
in
Fish."
Ph.
D.
thesis.
University
of
Utrecht,
Utrecht,
the
Netherlands.

H.
Konemann.
1981.
"
Quantitative
Structure­
Activity
Relationships
in
Fish
Toxicity
Studies.
Part
1:
Relationship
for
50
Industrial
Pollutants."
Toxicology,
19:
209­
221.

Konemann
H.
and
K.
Van
Leeuwen.
1980.
"
Toxicokinetics
in
Fish:
Accumulation
and
Elimination
of
Six
Chlorobenzenes
by
Guppies."
Chemosphere,
9:
3­
19.

Koss,
G.,
Seubert,
S.,
Seubert,
A.,
Seidel,
J.,
Koransky,
W.
and
H.
Ippen.
1983.
Studies
on
the
Toxicology
of
Hexachlorobenzene.
V.
Different
Phases
of
Porphyria
During
and
After
Treatment.
Archives
Toxicology,
52:
13­
22.
25
Koss,
G.,
Seubert,
S.,
Seubert,
A.,
Seidel,
J.,
Koransky,
W.
Ippen,
H.,
P.
Krauss.
1980.
Conversion
Products
of
Hexachlorobenzene
and
Their
Role
in
the
Disturbance
of
the
Porphyria
Pathway
in
Rats.
International
Journal
Biochemistry,
12(
5­
6):
1003­
1006.

Krishnan
K.,
Brodeur
J.,
G.
Plaa
et
al.
1992.
"
Modulation
of
Hexachlorobenzene­
induced
Hepatic
Porphyria
by
Methyl
Isobutyl
Ketone
in
the
Rat."
Toxicology
Letters,
61:
167­
174.

Lambrecht,
R.,
Ertrök,
E.,
Grunden,
E.
E.,
Peters,
H.
A.,
Morris,
C.
R.,
and
G.
T.
Bryan.
1983a.
Renal
Tumors
in
Rats
(
R)
Chronically
Exposed
to
Hexachlorobenzene
(
HCB).
Proceding
American
Association
Cancer
Research,
24:
59
(
Abstr).

Lambrecht,
R.,
Ertrök,
E.,
Grunden,
E.
E.,
Peters,
H.
A.,
Morris,
C.
R.,
and
G.
T.
Bryan.
1983b.
Hepatocarcinogenicity
of
Chronically
Administered
Hexachlorobenzene
in
Rats.
Federation
Proceeding,
42(
4):
786.

Laseter
J.,
Bartell
C.,
Laska
A.,
Holmquist,
D.,
Condie
D.,
Brown
J.
and
R.
Evans.
1976.
An
Ecological
Study
of
Hexachlorobenzene.
EPA­
560/
6­
76­
009.
National
Technical
Information
Service,
Springfield
VA.

Laska
A.,
Bartell
C.,
Condie
D.,
Brown
J.,
Evans
R.
and
J.
Laseter.
1978.
"
Acute
and
Chronic
Effects
of
Hexachlorobenzene
and
Hexachlorobutadiene
in
Red
Swamp
Crayfish
(
Procambarus
clarki)
and
Selected
Fish
Species.
Toxicology
and
Applied
Pharmacology,
43:
1­
12.

Mayer
F.
and
M.
Ellersieck.
1986.
Manual
of
Acute
Toxicity:
Interpretation
and
Data
Base
for
410
Chemicals
and
66
Species
of
Freshwater
Animals.
Resource
Publication
No.
160.
U.
S.
Fish
and
Wildlife
Service,
Washington
DC.,
pp.
265.

McLeese,
D.
W.
and
C.
D.
Metcalfe.
1980.
Toxicities
of
Eight
Organochlorine
Compounds
in
Sediment
and
Seawater
to
Crangon
Septemspinosa.
Bulletin
of
Environmental
Contamination
and
Toxicology,
25(
6):
921­
928.

Meller
M.,
Egeler
P.,
Rombke
J.,
Schallnass
H.,
Nagel
R.
and
B.
Streit.
1998.
"
Short­
term
Toxicity
of
Lindane,
Hexachlorobenzene,
and
Copper
sulfate
to
Tubificid
Sludgeworms
(
Oligochaeta)
in
Artificial
Media."
Ecotoxicology
and
Environmental
Safety,
39(
1):
10­
20.

Mortimer
M.
and
D.
Connell.
1995.
"
Effect
of
Exposure
to
Chlorobenzenes
on
Growth
Rates
of
the
Crab
Portunus
pelagicus
(
L)."
Environmental
Science
and
Technology,
28(
8):
1881­
1886.

Munoz
M.,
Ramos
C.,
and
J.
Tarazona.
1996.
"
Bioaccumulation
and
Toxicity
of
Hexachlorobenzene
in
Chlorella
vulgaris
and
Daphnia
magna."
Aquatic
Toxicology,
35:
211­
220.
26
National
Institute
for
Occupational
Safety
and
Health,
1983.
Registry
of
Toxic
Effects
of
Chemical
Substances,
U.
S.
Department
of
Health
and
Human
Services,
Public
Health
Service,
Centers
for
Disease
Control,
Cincinnati,
OH.

Nebeker
A.,
Griffis
W.,
Wise
C.,
Hopkins
E.,
and
J.
Barbitta.
1989.
"
Survival,
Reproduction
and
Bioconcentration
in
Invertebrates
and
Fish
Exposed
to
Hexachlorobenzene."
Environmental
Toxicology
and
Chemistry,
8:
601­
611.

Neuhauser,
E.,
Loehr,
R.
and
M.
Malecki.
1985.
"
Contact
and
Artificial
Soil
Tests
Using
Earthworms
to
Evaluate
the
Impact
of
Wastes
in
Soil,"
in
Petros,
J.,
Lacy
W.,
and
R.
Conway
(
Eds.),
Hazardous
and
Industrial
Solid
Waste
Testing:
4th
Symposium,
ASTM
STP
886,
Philadelphia,
PA,
886:
192­
203.

Niimi
A.,
and
C.
Cho.
1980.
"
Uptake
of
Hexachlorobenzene
(
HCB)
from
Feed
by
Rainbow
Trout
(
Salmo
gairdneri)."
Bulletin
of
Environmental
Contamination
and
Toxicology,
24:
834­
839.

Y.
Nishiuchi.
1980.
Toxicity
of
Formulated
Pesticides
to
Freshwater
Organisms.
LXXIV.
The
Aquiculture
 
Suisan
Zoshoku
28(
2):
107­
112
(
JPN).

Oliver
B.,
and
A.
Niimi.
1983.
"
Bioconcentration
of
Chlorobenzenes
from
Water
by
Rainbow
Trout:
Correlations
with
Partition
Coefficients
and
Environmental
Residues."
Environmental
Science
and
Technology,
17:
287­
291.

Parasher,
C.
D.,
Ozel,
M.
and
F.
Geike.
1978.
Effect
of
Hexachlorobenzene
and
Acetone
on
Algal
Growth:
Physiology
and
Ultrastructure.
Chemical
Biological
Interacttion,
20(
1):
89­
95.

Parrish,
P.
R.,
Cook,
G.
H.
and
J.
M.
Patrick,
Jr.
1974.
Hexachlorobenzene:
Effects
on
Several
Estuarine
Animals.
In:
Proc.
28th
Annual
Conf.
erence
S.
E.
Association
Game
Fish
Commision,
pp.
179­
187.

Rizzardini,
MK.
And
A.
G.
Smith.
1982.
Sex
Differences
in
the
Metabolism
of
Hexachlorobenzene
by
Rats
and
the
Development
of
Porphyria
in
Females.
Biochemistry
Pharmacology,
31(
22):
3543­
3548.

Rush,
G.
F.,
Smith,
J.
H.
Maita,
K.
et
al.
1983.
Perinatal
Hexachlorobenzene
Toxicity
in
the
Mink.
Environmental
Research,
31:
116­
124.

Sabourin
T.,
Faulk
R.,
Coyle
J.,
DeGraeve
G.,
Brooke
L.,
Call
D.,
Harting
S.,
Larson
L.,
Lindbergh
C.,
Markee
T.,
McCauley
D.,
and
S.
Poirier.
1986.
Freshwater
Aquatic
Citeria
Development
and
Testing.
Final
Report
by
Battelle
to
U.
S.
EPA
on
Contract
No.
68­
01­
6986.
Office
of
Water
Regulations
and
Standards,
Criteria
and
Standards
Div.,
Washington,
DC.
27
Sample,
B.
E.,
Opresko,
D.
M.
and
G.
W.
Suter
II.
1996.
Toxicological
Benchmarks
for
Wildlife:
1996
Revision.
U.
S.
Department
of
Energy,
Office
of
Environmental
Management,
Oak
Ridge.
ES/
ER/
TM­
86/
R3.

Savitskii
IV.
1964.
The
Basis
for
Determining
Safe
Permissible
Concentrations
of
Hexachlorobenzene
and
Pentachloronitrobenzene
in
the
Air.
Vopr
Prom
I
Sel'skokhoz
Toksikol:
158­
173.

Savitskii
IV.
1965.
The
Basis
for
Determining
Safe
Permissible
Concentrations
of
Hexachlorobenzene
and
Pentachloronitrobenzene
in
the
Air.
Chemical
Abstracts,
63:
8952.

Schafer,
E.
and
W.
Bowles.
1985.
"
Acute
Oral
Toxicity
and
Repellency
of
933
Chemicals
to
House
and
Deer
Mice,"
Archives
of
Environmental
Contamination
and
Toxicology,
14:
111­
129.

Smith
A.,
Cabral
J.,
Carthew
P.
et
al.
1989.
"
Carcinogenicity
of
Iron
in
Conjunctionn
with
a
Chlorinated
Environmental
Chemical,
Hexachlorobenzene,
in
C57BL/
1OscSn
mice.
International
Journal
Cancer,
43:
492­
496.

Smith
A.,
Francis
J.,
Dinsdale
D.
et
al.
1985.
"
Hepatocarcinogenicity
of
Hexachlorobenzene
in
Rats
and
the
Sex
Difference
in
Hepatic
Iron
Status
and
Development
of
Porphyria."
Carcinogenesis,
6(
4):
631­
636.

Smith,
E.
N.
and
G.
P.
Carlson.
1980.
Various
Pharmacokinetic
Parameters
in
Relation
to
Enzyme­
inducing
Abilities
of
1,3,4­
trichlorobenzene
and
1,2,4­
tribromobenzene.
J.
Toxicology
Environmental
Health,
6(
4):
737­
749.

Smith,
A.
G.,
Francis,
J.
E.
and
F.
De
Matteis.
1980.
Lobes
of
Rat
Tissue
Respond
at
Different
Rates
to
Challenge
by
Dietary
Hexachlorobenzene.
Biochemistry
Pharmacology,
29(
23):
3127­
3131.

Spehar,
R.
1986.
U.
S.
EPA,
Duluth,
MN.
(
Memorandum
to
D.
J.
Call,
University
of
Wisconsin­
Superior,
Superior,
WI,
September
16).

Sugatt
R.,
O'Grady
D.
and
S.
Banerjee.
1984.
"
Toxicity
of
Organic
Mixtures
Saturated
in
Water
to
Daphnia
magna;
Effect
of
Compositional
Changes."
Chemosphere,
13:
11­
18.

Sundlof,
S.
M.,
Parker,
A.
J.,
Simon,
J.,
Dorner,
J.
L.
and
L.
G.
Hansen.
1981.
Sub­
acute
Toxicity
of
Hexachlorobenzene
in
Female
Beagles,
Including
Electroencephalographic
Changes.
Veterinary
Human
Toxicology,
23(
2):
92­
96.
28
U.
S.
Environmental
Protection
Agency
(
U.
S.
EPA).
1985.
Health
Assessment
Document
for
Chlorinated
Benzenes.
Office
of
Health
and
Environmental
Assessment,
Washington
DC,
EPA/
600/
8­
84/
015F.

U.
S.
Environmental
Protection
Agency
(
U.
S.
EPA).
1988.
Ambient
Aquatic
Life
Water
Quality
Criteria
for
Hexachlorobenzene.
Office
of
Research
and
Development
 
Environmental
Research
Laboratory,
Duluth
Minnesota.
Doc.
No.
440­
5­
88­
092.

U.
S.
Environmental
Protection
Agency
(
USEPA).
1989.
Risk
Assessment
Guidance
for
Superfund.
Volume
II:
Environmental
Evaluation
Manual.
Interim
Final.
EPA/
540/
1­
89/
001.
March,
1989.

U.
S.
Environmental
Protection
Agency
(
USEPA).
1993.
Wildlife
Exposure
Factors
Handbook.
Volume
I.
Office
of
Research
and
Development,
Washington,
D.
C.
EPA/
630/
R­
93/
187a.

U.
S.
Environmental
Protection
Agency
(
U.
S.
EPA).
2001.
Ecotox
Database.

U.
S.
National
Library
of
Medicine.
Hazardous
Substances
DataBank
(
HSDB).
1995.
Bethesda,
MD,
6­
18.

Veith
G.,
Call
D.
and
L.
Brooke.
1983a.
Estimating
the
Acute
Toxicity
of
Narcotic
Industrial
Chemicals
to
Fathead
Minnows.
In:
Aquatic
Toxicology
and
Hazard
Assessment:
Sixth
Symposium.
Bishop
W.,
Cardwell
R.
and
B.
Heidolph,
(
Eds.).
ASTM
STP
802.
American
Society
for
Testing
and
Materials,
Philadelphia
PA,
pp.
90­
97.

Veith
G.,
Call
D.
and
L.
Brooke.
1983b.
Structure­
toxicity
Relationships
for
the
Fathead
Minnow,
Pimephales
promelas:
Narcotic
Industrial
Chemicals.
Canadian
J.
Fisheries
and
Aquatic
Science,
40:
743­
748.

Veith
G.,
DeFoe,
D.,
and
B.
Bergstedt.
1979.
"
Measuring
and
Estimating
the
Bioconcentration
Factor
of
Chemicals
in
Fish."
Journal
Fisheries
Research
Board
Canadian,
36:
1040­
1048.

Vincent
S.,
Smith
A.
and
U.
Muller­
Eberhard.
1989.
"
Modulation
of
Hepatic
Heme­
binding
Z
Protein
in
Mice
by
the
Porphyrogenic
Carcinogens
Griseofulvin
and
Hexachlorobenzene."
Cancer
Letters,
45:
109­
114.

Vos,
J.,
Van
der
Maas,
H.,
Musch,
A.,
and
E.
Ram.
1971.
"
Toxicity
of
Hexachlorobenzene
in
Japanese
Quail
with
Special
Reference
to
Porphyria,
Liver
Damage,
Reproduction
and
Tissue
Residues,"
Toxicology
and
Applied
Pharmacology,
18:
944­
957.

Weseloh,
D.,
Teeple,
S.
and
M.
Gilbertson.
1983.
Double­
breasted
Comorants
of
the
Great
Lakes;
egg­
laying
parameters,
reproductive
failure,
and
contaminant
residues
in
eggs,
Lake
Huron
1972­
1973.
Canadian
Journal
of
Zoology,
61:
427­
436.
29
Wong
P.,
Chau
Y.,
Rhamey
J.
and
M.
Docker.
1984.
"
Relationship
Between
Water
Solubility
of
Chlorobenzenes
and
their
Effects
on
a
Freshwater
Green
Alga."
Chemosphere,
13:
991­
996.

Yoshioka
Y.,
Ose
Y.
and
T.
Sato.
1986.
"
Correlation
of
the
Five
Test
Methods
to
Assess
Chemical
Toxicity
and
Relation
to
Physical
Properties."
Ecotoxicology
Environmental
Safety,
12:
15­
21.

G.
Zaroogian.
1981.
Interlaboratory
Comparison
 
Acute
Toxicity
Tests
Using
the
48
Hour
Oyster
Embryo­
Larval
Assay.
U.
S.
EPA,
Narragansett,
RI:
17.
30
Appendix
1.
Toxicity
Data
for
Hexachlorobenzene
(
EPA,
1985
 
Health
Assessment
for
Chlorinated
Benzenes)

Species
Exposure
Duration/
Type
Endpoint
­
Concentration
Reference
Fish
Acute/
Subchronic
Toxicity
Data
Largemouth
black
bass
(
Micropterus
salmoides)
10
days
(
static)

15
days
(
static)
NOAEL
10

g/
L
NOAEL
26

g/
L
Laska
et
al.,
1978
Laska
et
al.,
1978
Sheepshead
minnow
(
C.
variegatus)
96
hrs.
(
constant­
flow)
NOAEL
0.13
mg/
L
Parrish
et
al.,
1974
(
conc
were
nominal
actual
conc
were
probably
lower)
Pinfish
(
Lagodon
rhomboides)
96
hrs.
(
constant­
flow)
NOAEL
1.0
mg/
L
Parrish
et
al.,
1974
(
conc
were
nominal
actual
conc
were
probably
lower)
Rainbow
trout
(
S.
gairdneri)
48
hrs.
LC50
>
0.03
mg/
La
Calamari
et
al.,
1983
Brachydanio
rerio
48
hrs.
LC50
>
0.03
mg/
La
Calamari
et
al.,
1983
Aquatic
Crustaceans
 
Acute
Toxicity
Data
Crayfish
(
Procambarus
clarkii)
Duration
unspecified
(
static
or
flow­
through
system)
0.02
mg/
L
 
NOAEL
Laska
et
al.,
1978
(
saturated
aqueous
solution)
Shrimp
(
Crangon
septemspinosa)
96
hr.
(
static)
0.0072
(
mg/
L)
 
no
mortality
McLeese
and
Metcalfe,
1980
Water
flea
(
Daphnia
magna)
24
hr.
IC50
>
0.03
mg/
La
(
immobilization
concentration
for
50
percent
of
animals)
Calamari
et
al.,
1983
Aquatic
Plants
 
Acute
Toxicity
Data
Freshwater
algae
(
Selenastrum
capricornatum)
96
hr.
(
static)
EC50
>
0.03
mg/
L
Calamari
et
al.,
1983
Mixed
diatoms
and
green
algae
(
Thalassiosira
pseudonana,
Dumaliella
tertiolecta)
72
hrs.
(
static)
0.1
mg/
L
NOEC
 
no
growth
inhibition
Biggs
et
al.,
1979
Green
algae
(
Chlorella
pyrenoidosa)
76
hrs.
10.0
mg/
L
 
Growth
reduction
(
growth
reduced
to
87.5
percent
of
control
cultures)
measured
by
dry
mass
Parasher
et
al.,
1978
a
Although
the
endpoint
values
numbers
were
reported
in
some
sources
as
less
than
("<")
the
specific
test
concentration,
according
to
the
text
of
the
actual
Calamari
et
al.
(
1983)
study,
it
appears
that
the
endpoints
were
greater
than
the
actual
concentration
(
e.
g.,
>
0.03
mg/
L)
31
Appendix
2.
Toxicity
Values
for
Aquatic
Animals
and
Plants
(
EPA,
1988)

Species
Test
Method
/
Durationa
LC50
or
EC50
(

g/
L)
b
Reference
Acute/
Subchronic
Toxicity
Hydra
(
adult)
static,
measured/
(
duration
not
reported)
>
65.9
Sabourin
et
al.
1986
Leech
(
adult)
(
Nephelopsis
obscura)
static,
measured/
(
duration
not
reported)
>
73.2
Sabourin
et
al.
1986
Snail
(
adult)
(
Aplexa
hypnorum)
static,
measured/
(
duration
not
reported)
>
77.6
Sabourin
et
al.
1986
Cladoceran
(
Ceriodaphnia
dubia)
Static,
measured/
(
7
days)
EC50
>
7.0
Spehar
1986
Cladoceran
<
24
hr.
(
Daphnia
magna)
28.3
hr
>
6
(
EC10)
(
immobilization)
Sugatt
et
al.,
1984
Cladoceran
~
5
days
(
Moina
macrocopa)
3
hr
>
6
(
LC50)
Yoshioka
et
al.,
1986
Amphipod
(
adult)
(
Gammarus
pseudolimnaeus)
Static,
measured/
(
duration
not
reported)
>
77.6
Sabourin
et
al.
1986
Crayfish
(
adult)
(
Procambarus
clarki)
Flow­
through,
measured/
(
10
days)
LC50
>
27.3c
Laseter
et
al.,
1976
Laska
et
al.,
1978
Crayfish
(
juvenile)
(
Procambarus
sp.)
Static,
measured/
(
duration
not
reported)
LC50
>
5.2d
Laseter
et
al.,
1976
Laska
et
al.,
1978
Stonefly
(
nymph)
(
Pteronarcys
sp.)
Static,
measured/
(
duration
not
reported)
>
69.7
Sabourin
et
al.
1986
Midge
(
3rd­
4th
instar),
(
Tanytarsys
dissimilis)
Static,
measured/
(
duration
not
reported)
>
58.1
Call
et
al.
1983
Rainbow
Trout
(
juvenile)
(
Salmo
gairdneri)
Flow­
through,
measured/
(
duration
not
reported)
>
3.76
Spehar,
1986
Rainbow
trout
(
Salmo
gairdneri)
24
hr
LC50
>
30
Calamari
et
al.,
1983
Fathead
minnow
(
Pimephales
promelas)
Flow­
through,
measured/
(
duration
not
LC50
>
6e
Veith
et
al.,
1983a,
b
Ahmad
at
al.,
1984
Carlson
and
Kosian,
1987
Species
Test
Method
/
Durationa
LC50
or
EC50
(

g/
L)
b
Reference
32
reported)
Fathead
minnow
(
0.7
g)
(
Pimephales
promelas)
Static,
unmeasured
(
duration
not
reported)
LC50
>
10,000
Johnson
and
Finley
1980,
Mayer
and
Ellersieck
1986
Appendix
2.
Toxicity
Values
for
Aquatic
Animals
and
Plants
(
EPA,
1988)
(
Cont'd)

Species
Test
Method
/
Durationa
LC50
or
EC50
(

g/
L)
b
Reference
33
Channel
catfish
(
0.8g)
(
Ictalurus
punctatus)
Static,
unmeasured
(
24
hrs.)
LC50
­
13,500
Johnson
and
Finley
1980,
Mayer
and
Ellersieck
1986
Channel
catfish
(
1.3g)
(
Ictalurus
punctatus)
Static,
unmeasured
(
duration
not
reported)
LC50
>
100,000
Mayer
and
Ellersieck
1986
Channel
catfish
(
sac
fry)
(
Ictalurus
punctatus)
Static,
unmeasured
(
24
hrs.)
LC50
­
7,000
Mayer
and
Ellersieck
1986
Bluegill
(
1.5g)
(
Lepomis
macrochirus)
Flow­
through,
measured/
(
duration
not
reported)
>
78.4
Call
et.
al,
1983
Bluegill
(
1g)
(
Lepomis
macrochirus)
Static,
unmeasured
(
duration
not
reported)
LC50
­
12,000
Johnson
and
Finley
1980,
Mayer
and
Ellersieck
1986
Bluegill
(
1.6g)
(
Lepomis
macrochirus)
Flow­
through,
unmeasured
(
duration
not
reported)
LC50
>
1,000
Mayer
and
Ellersieck
1986
Largemouth
bass
(
Micropterus
salmoides)
Flow­
through,
measured/
(
10
days)
LC50
>
25.8e
Laseter
et
al.,
1976
Laska
et
al.,
1978
Largemouth
bass
(
0.5g)
(
Micropterus
salmoides)
Static,
unmeasured
 
96
hrs.
LC50
­
12,000
Johnson
and
Finley
1980,
Mayer
and
Ellersieck
1986
Largemouth
bass
10
days
3.5

g/
L
(
liver
and
kidney
damage)
Laseter
et
al.,
1976
Coho
salmon
Static
­
96
hrs.
LC50
­
50,000
Johnson
and
Finley,
1980
Guppy
(
2­
3
mo.)
(
Poecilia
reticulata)
14
days
>
320
(
LC50)
Konemann
1979,
1981
Green
alga
(
Ankistrodesmus
falcatus)
4
hr
>
3
(
EC50)
Wong
et
al.,
1984
Protozoan
(
Colpidium
compylum)
43
hr
10,000
(
No
effect)
Dive
et.
al.,
1980
Protozoan
(
Tetrahymena
pyriformis)
24
hr
>
6
(
EC50)
Yoshioka
et
al.,
1985
Cladoceran
(
Daphnia
magna)
24
hr
>
30
(
EC50)
Calamari
et
al.,
1983
Alga,
(
Scenedesmus
subspicatus)
4
day
duration
>
10
(
EC10)
Geyer
et
al.,
1985
Alga,
(
Scenedesmus
subspicatus)
4
day
duration
>
10
(
EC50)
Geyer
et
al.,
1985
Appendix
2.
Toxicity
Values
for
Aquatic
Animals
and
Plants
(
EPA,
1988)
(
Cont'd)

Species
Test
Method
/
Durationa
LC50
or
EC50
(

g/
L)
b
Reference
34
Alga,
(
Selenastrum
capricornatum)
4
day
duration
>
30
(
EC50)
Calamari,
1983
CHRONIC
TOXICITY
Cladoceran,
(
Ceriodaphnia
dubia)
Life
cycle/
partial
life
cycle
 
7
days
>
7f
(
NOEC)
Spehar,
1986
Rainbow
trout
(
Salmo
gairdneri)
Early
life
stage
 
90
days
>
3.68f
(
NOEC)
Spehar,
1986
Fathead
minnow
(
Pimephales
promelas)
Early
life
stage
 
32
days
>
4.8f
(
NOEC)
Ahmad
et
al.,
1984:
Carlson
and
Kosian,
1987
Cladoceran
(
Daphnia
magna)
14
days
16
(
50
percent
reduction
in
fertility)
 
EC50
Calamari
et
al.,
1983
a
For
several
of
the
acute
and
subchronic
toxicity
data,
the
test
duration
was
not
reported.
b
For
several
of
the
acute
and
subchronic
toxicity
data,
the
distinction
was
not
made
between
the
LC50/
EC50
endpoint.
c
The
exposure
was
continued
for
a
period
of
ten
days
without
any
mortalities
attributable
to
hexachlorobenzene.
d
The
exposure
was
continued
for
a
period
of
8
days
without
any
mortalities
attributable
to
hexachlorobenzene.
e
The
LC50
was
above
saturation,
which
was
estimated
to
be
6

g/
L
based
on
Metcalf
et
al.,
1973.
f
Highest
tested
concentration;
no
tested
concentration
caused
an
unacceptable
effect.
35
Appendix
3.
Aquatic
Toxicity
Data
from
ECOTOX
Database
(
EPA,
2001)

Species
Duration/
Type
Endpoint/
Concentration
Reference
Acute/
Subchronic
Toxicity
Channel
catfish
(
Ictalurus
punctatus)
24
hrs.
exp
type
 
static
LC50
 
11.3,
16.2
mg/
L
Mayer
and
Ellersieck,
1986
Channel
catfish
(
Ictalurus
punctatus)
24
hrs.
exp
type
 
static
LC50
>
100
mg/
L
Mayer
and
Ellersieck,
1986
Channel
catfish
(
Ictalurus
punctatus)
24
hrs.
exp
type
 
static
LC50
 
7.0,
6.4,
7.7
mg/
L
Mayer
and
Ellersieck,
1986
Channel
catfish
(
Ictalurus
punctatus)
96
hrs.
exp
type
­
static
LC50
 
14,
11,16
mg/
L
Johnson
W.
and
M.
Finley,
1980
Brine
shrimp
(
Artemia
salina)
24
hrs.
exp
type
 
static
LC50
>
0.0116
mmol/
m3
Abernathy,
S.,
Bobra,
A.,
Shiu
W.,
Wells
P.,
and
D
MacKay,
1986
Toad
(
Bufo
bufo
japonicus)
24
hrs.
exp
type
 
static
LC50
>
4.2
mg/
L
Nishiuchi,
Y.
1980
American/
virginia
oyster
(
Crassostrea
virginica)
48
hrs.
exp
type
 
not
reported
EC50
 
6.5
mg/
L
(
morphology)
Zaroogian,
G.,
1981
Diatom
(
Cyclotella
meneghiniana)
48
hrs.
exp
type
 
static
EC50
­
2

g/
L
(
genetic
effect)
Figueroa,
I.
And
M.
Simmons,
1991
Zebra
danio
(
Danio
rerio)
48
hrs.
exp
type
 
static
LC50
>
30

g/
L
Calamari,
D.,
Galassi,
S.,
Setti,
F.,
and
Vighi,
M.,
1983
Water
flea
(
Daphnia
magna)
48
hrs.
exp
type
 
static
EC50
>
0.0166
mmol/
m3
(
intoxication)
Abernathy,
S.,
Bobra,
A.,
Shiu
W.,
Wells
P.,
and
D
MacKay,
1986
Water
flea
(
Daphnia
magna)
14
days
Exp
Type
­
Renewal
EC50
 
16

g/
L
(
reproduction)
Calamari,
D.,
Galassi,
S.,
Setti,
F.,
and
Vighi,
M.,
1983
Water
flea
(
Daphnia
magna)
21
days
Exp
Type
­
Renewal
NOEC
>
37.9
(

g/
L)
(
behavioral
effects)
Caspers,
N.,
Hartmann,
P.,
Kanne,
R.,
and
Knoop,
G.,
1993.
Water
flea
(
Daphnia
magna)
21
days
Exp
Type
­
Renewal
NOEC
>
6.7
(

g/
L)
(
reproduction
effects)
Caspers,
N.,
Hartmann,
P.,
Kanne,
R.,
and
Knoop,
G.,
1993.
Green
algae
(
Haematococcus
pluvialis)
Duration
and
exposure
type
not
reported
EC10
>
40

g/
L
(
population
effect)
Knie,
J.,
Halke
A.,
Juhnke,
I.,
and
W.
Schiller,
1983
Channel
catfish
(
Ictalurus
punctatus)
24
h
exp
type
 
static
LC50
 
13.5,
11.3,
16.2
mg/
L
Mayer
and
Ellersieck,
1986
Appendix
3.
Aquatic
Toxicity
Data
from
ECOTOX
Database
(
EPA,
2001)
(
Cont'd)

Species
Duration/
Type
Endpoint/
Concentration
Reference
36
Channel
catfish
(
Ictalurus
punctatus)
24
hrs.
exp
type
 
static
LC50
>
100
mg/
L
Mayer
and
Ellersieck,
1986
Channel
catfish
(
Ictalurus
punctatus)
24
hrs.
exp
type
 
static
LC50
 
7.0,
6.4,
7.7
mg/
L
Mayer
and
Ellersieck,
1986
Channel
catfish
(
Ictalurus
punctatus)
96
hrs.
exp
type
 
static
LC50
 
14,
11,16
mg/
L
Johnson
W.
and
M.
Finley,
1980
37
Appendix
4.
Additional
Aquatic
Toxicity
Data
­
Miscellaneous
Sources
Species
Exposure
Duration/
Type
Endpoint/
Concentration
Source
Amphipod
(
Hyalella
azteca)
14
days
(
spiked
sediment)
NOEC
 
42
mg/
kg
Barber
et
al.,
1997
Midge
(
Chironomus
tentans)
14
days
(
spiked
sediment)
NOEC
 
42
mg/
kg
Barber
et
al.,
1997
Estuarine
amphipod
(
Leptocheirus
plumulosus)
10
days
(
spiked
sediment)
NOEC
 
60
mg/
kg
Fuchsman
et
al.,
1998
Freshwater
amphipod
(
Hyalella
azteca)
10
days
(
spiked
sediment)
NOEC
 
60
mg/
kg
Fuchsman
et
al.,
1998
Midge
(
Chironomus
tentans)
10
days
(
spiked
sediment)
NOEC
 
60
mg/
kg
Fuchsman
et
al.,
1998
tubificid
sludgeworms
(
oligochaeta)
72
hrs.
LC50/
EC50
>
1000
mg/
kg
Meller
et
al.,
1998
Flounder
(
Platichthys
flesus)
96
hr
LC50
 
0.7
uM
(
equivalent
to
0.2
mg/
L)
Furay
and
Smith,
1995
Sole
(
Solea
solea)
96
hr
LC50
 
0.5
uM
(
equivalent
to
0.14
mg/
L)
Furay
and
Smith,
1995
Water
flea
(
Daphnia
magna)
48
hr
(
flow
through)
LC50
­
>
5.0

g/
L
Nebeker
et
al.,
1989
(
no
significant
mortality
observed
at
highest
conc.
tested
 
saturation
snail
(
Lymnaea
palustris)
10­
12
weeks
0.5

g/
L,
5

g/
L
(
significantly
reduced
growth
rate)
Baturo
et
al.,
1995
Crab
(
Portunus
pelagicus)
6
weeks
5

g/
L
(
reduction
in
carapace
growth
rate)
Mortimer
and
Connell,
1995
Freshwater
amphipod
(
Hyalella
azteca)
30
days
(
flow
through)
(
higher
HCB
water
conc.
of
2
separate
trials)
NOEC
 
4.5

g/
L
Mean
adult
tissue
value
of
172

g/
g
Nebeker
et
al.,
1989
(
no
significant
mortality
or
effects
to
reproduction
and
growth
observed
at
near
water
saturation
Fathead
minnow
(
Pimephales
promelas)
28
days
(
flow
through)
NOEC
­
3.8

g/
L
Mean
tissue
value
of
47

g/
g
Nebeker
et
al.,
1989
(
no
mortality
or
difference
in
growth
rate
observed).

Worm
(
Lumbriculus
variegatus)
49
days
(
flow
through)
NOEC
 
4.7

g/
L
Mean
tissue
value
of
223

g/
g
Nebeker
et
al.,
1989
(
no
significant
effects
on
survival,
growth
or
asexual
reproduction
were
observed
38
Appendix
5.
Terrestrial
Toxicity
Data
(
Invertebrates,
Birds,
Mammals)
 
Ecotox
Database,
EPA
2001
Species
Test
Duration/
Typea
Endpoint/
Concentration
Reference
Terrestrial
Invertebrates
Earthworm
(
Eisenia
fetida)
48
hrs
(
dosed
one
time
during
the
study)
LC50
>
1000

g/
cm2
Neuhauser,
1985
Earthworm
(
Eisenia
fetida)
28
days
LC50
>
1000
mg/
kg
Ballhorn,
1984
Birds
Mallard
Duck
(
Anas
platyrhynchos)
8
days
LC50
>
5000
ppm
Hill
et
al.,
1975
Ring­
necked
pheasant
(
Phasianus
colchicus)
8
days
(
exposed
to
HCB
for
a
duration
of
5
days)
LC50
 
617
ppm
Hill
et
al.,
1975
Japanese
quail
(
Coturnix
japonica)
5
days
(
dose
admin
1x
per
day
for
5
days)
LOEL
 
500
mg/
kg
(
Biochemistry;
microsomal
protein
response
and
enzymes;
7­
ethoxyresorufin
O­
deethylase
and
glutathione
S­
transferase
level)
Carpenter,
1985a
Japanese
quail
(
Coturnix
japonica)
5
days
(
dose
admin
1x
per
day
for
5
days)
NOEL
 
500
mg/
kg
(
Growth:
whole
body
weight,
and
morphology:
organ
weight
related
to
body
weight
Carpenter,
1985a
Japanese
quail
(
Coturnix
japonica)
5
days
(
dose
admin
1x
per
day
for
5
days)
NOEL
 
500
mg/
kg
(
Enzymes,
epoxide
hydrase,
NADPH
cytochrome
C
reductase,
and
cytochrome
P­
450
levels)
Carpenter,
1985a
Japanese
quail
(
Coturnix
japonica)
15
days
(
dose
admin
1x
per
day
for
test
duration)
LOEL
 
100
mg/
kg
(
biochemistry,
protoporphyrin
level)
Carpenter,
1985b
Japanese
quail
(
Coturnix
japonica)
17
days
(
dose
admin
1x
per
day
for
test
duration)
LOEL
 
100
mg/
kg
(
Enzymes,
ALA
synthetase
level)
Carpenter,
1985b
Japanese
quail
(
Coturnix
japonica)
17
days
(
dose
admin
1x
per
day
for
test
duration)
LOEL
 
100
mg/
kg
(
Morphology,
organ
weight
related
to
body
weight)
Carpenter,
1985b
Japanese
quail
(
Coturnix
japonica)
90
days
LOEL
 
20
ppm
(
Egg
volume)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
90
days
NOEL
 
5
ppm
(
Egg
volume)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
90
days
LOEL
 
5
ppm
(
Morphology,
organ
weight
related
to
body
weight)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
90
days
NOEL
 
1
ppm
(
Morphology,
organ
weight
related
to
body
weight)
Vos,
1971
Appendix
5.
Terrestrial
Toxicity
Data
(
Invertebrates,
Birds,
Mammals)
 
Ecotox
Database,
EPA
2001
(
Cont'd)

Species
Test
Duration/
Typea
Endpoint/
Concentration
Reference
39
Japanese
quail
(
Coturnix
japonica)
90
days
LOEL
 
5
ppm
(
Morphology,
organ
weight)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
90
days
NOEL
 
1
ppm
(
Morphology,
organ
weight)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
4
weeks
LOEL
 
120
ppm
(
Biochemistry,
hemoglobin
level
and
packed
cell
volume)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
4
weeks
NOEL
 
20
ppm
(
Biochemistry,
hemoglobin
level
and
packed
cell
volume)
Vos,
1971
Japanese
quail
(
Coturnix
japonica)
4
weeks
LOEL
 
20
ppm
(
Enzymes,
glutamicoxaloacetic
transaminase)
Vos,
1971
Mammals
Deer
mouse
(
Peromyscus
maniculatus)
3
days
(
dosed
one
time
per
study
period)
LD50
 
710
mg/
kg
Schafer
and
Bowles,
1985
a
All
tests
are
oral
unless
otherwise
indicated.
40
Appendix
6.
Terrestrial
Toxicity
Data
­
EPA
1985,
Health
Assessment
for
Chlorobenzenes
(
Mammals
Only)

Species
Test
Duration/
Type
Endpoint/
Concentration
Reference
Intermediate/
Subchronic
Exposure
Rat
43
days
(
dosed
every
other
day)
100
mg/
kg
 
effect:
suggested
covalent
bonding
of
HCB
metabolites
to
cytosolic
proteins.
Koss
et
al.,
1980a
Rat
98
days
100
mg/
kg
diet
 
effect:
porphyria
(
increased
liver
lobe
porphyrins),
decreased
activity
of
uroporphyrinogen
decarboxylase
Smith
et
al.,
1980
Rat
103
days
(
dosed
every
other
day)
14
mg/
kg
 
effect:
porphyria
in
treated
females,
poss.
related
to
estrogen
levels
Rizzardini
and
Smith,
1982
Rat
(
females)
6
weeks
exposed
(
dosed
every
other
day)
then
held
for
additional
18
months
100
mg/
kg
 
effect:
porphyria
(
liver
uroporphyrin
levels
peaked
7
months
postexposure
hadn't
returned
to
normal
by
18
months)
Koss
et
al.,
1983
Dog
(
female)
21
days
50
or
150
mg/
kg/
day
 
liver
and
hepatocyte
enlargement,
electroencephalogram
dysrhythmias
Sundlof
et
al.,
1981
Monkey
(
nursing)
60
days
7.5
 
186
ppm
milk
 
2
of
3
infants
died
as
a
result
of
exposure
Bailey
et
al.,
1980
Mink
From
gestation
until
17
weeks
of
age
1
or
5
mg/
kg
diet
 
dose­
related
increase
in
offspring
mortality,
induction
of
hepatic
enzymes
in
exposed
offspring.
Rush
et
al.,
1983
Hamster
90
days
200
mg/
kg
diet
 
hepatic
lesions,
bile
duct
hyperplasias
and
hepatomas
Lambrecht
et
al.,
1982
Appendix
6.
Terrestrial
Toxicity
Data
­
EPA
1985,
Health
Assessment
for
Chlorobenzenes
(
Mammals
Only)
(
Cont'd)

Species
Test
Duration/
Type
Endpoint/
Concentration
Reference
41
Chronic
Exposure
Rat
(
females)
75­
90
weeks
6­
8
mg/
kg/
day
 
effect:
decline
in
body
weight,
porphyria,
enlarged
liver
and
liver
tumors
Smith
and
Cabral,
1980
Rat
Up
to
2
years
4­
5
mg/
kg/
day
 
effect:
porphyria,
increased
heptomas,
heptocarcinomas,
bile
duct
adenomas,
renal
adenomas
and
renal
carcinomas
Lambrecht
et
al.,
1983,
a,
b
Rat
Gestation
through
lifetime
(
130
weeks)
0.32
mg/
kg
diet
­
NOEC
Arnold
et
al.,
1985
Rat
Gestation
through
lifetime
(
130
weeks)
 
chronic
1.6
mg/
kg
diet
 
NOAEL
for
liver
effects
(
0.08
mg/
kg/
day)
Arnold
et
al.,
1985
Rat
Gestation
through
lifetime
(
130
weeks)
 
chronic
8.0
mg/
kg
diet
 
LOAEL
­
increase
in
liver
pathologies
(
0.29
mg/
kg/
day)
Arnold
et
al.,
1985
Rat
Gestation
through
lifetime
(
130
weeks)
 
chronic
40
mg/
kg
diet
 
increased
mortality
as
pups,
increase
in
liver
and
kidney
pathologies,
and
parathyroid
tumors
in
males
Arnold
et
al.,
1985
Rat
Fo
to
F1a
and
F1b
generations
21­
day
LD­
50
values
for
pups
were
100
and
140
mg/
kg
for
the
F1a
and
F1b
generations,
respectively
Kitchin
et
al.,
1982
42
Appendix
7.
Terrestrial
Toxicity
Data
(
Mammals)
­
ATSDR,
2000
 
Toxicity
Profile
for
Hexachlorobenzene
Species
Test
Duration/
Type
Endpoint/
Concentration
Reference
Acute
Exposure
Rat
NRa
LD50
 
3,500
mg/
kg
Savitskii
(
1964,
1965)
Mouse
NR
LD50
 
4,000
mg/
kg
Savitskii
(
1964,
1965)
Rabbit
NR
LD50
 
2,600
mg/
kg
Savitskii
(
1964,
1965)
Cat
NR
LD50
 
1,700
mg/
kg
Savitskii
(
1964,
1965)
Intermediate/
Subchronic
Exposure
Rat
80
days
100
mg/
kg/
day
 
lethality
(
60
of
90
individuals
died)
Cuomo
et
al.,
1991
Rat
13
weeks
19
mg/
kg/
day
 
lethality
(
4
of
9
individuals
died)
Den
Besten
et
al.
(
1993)

Rat
80
days
LOAEL
 
100
mg/
kg/
day
(
increased
hepatic
porphyrins)
Cantoni
et
al.
(
1990)

Rat
8
weeks
(
1x
per
day)
LOAEL
 
1000
mg/
kg/
day
(
increased
liver
weight
and
and
hepatic
porphyrins)
Kleiman
de
Pisarev
et
al.
(
1990)

Rat
4
weeks
(
1x
per
day)
LOAEL
 
1000
mg/
kg/
day
(
increased
liver
weight)
Kleiman
de
Pisarev
et
al.
(
1995)

Rat
30
days
(
dosed
one
time
per
day)
LOAEL
 
500
mg/
kg/
day
(
endocrine
system,
decreased
serum
T4
levels)
Kleiman
de
Pisarev
et
al.
1989
Rat
6
weeks
(
1
time/
day,
5
days
per
week)
LOAEL
 
50
mg/
kg
(
increased
hepatic
and
urinary
porphyrins
Krishnan
et
al.
(
1992)

Rat
56
days
LOAEL
 
5
mg/
kg/
day
(
increased
highly
carboxylated
porphyrins
in
the
liver)
Kennedy
and
Wigtield
(
1990)

Rat
5
weeks,
(
dosed
1x
per
day,
5
days
per
week)
LOAEL
 
1
mg/
kg/
day
(
increased
liver
weight)
Andrews
et
al
(
1988)

Mouse
7
weeks
LOAEL
 
26
mg/
kg/
day
(
increased
hepatic
porphyrins)
Vincent
et
al.
(
1989)

Chronic
Exposure
Rat
2
generations
NOAEL
 
2
mg/
kg/
day
(
hematological,
renal
and
body
weight
effects):
Arnold
et
al.
(
1985)
Appendix
7.
Terrestrial
Toxicity
Data
(
Mammals)
­
ATSDR,
2000
 
Toxicity
Profile
for
Hexachlorobenzene
(
Cont'd)

Species
Test
Duration/
Type
Endpoint/
Concentration
Reference
43
Rat
2
generations
NOAEL
 
0.08
mg/
kg/
day
(
hepatic
effects)
Arnold
et
al.
(
1985)

Rat
2
generations
LOAEL
 
0.4
mg/
kg/
day
(
increased
liver
weight)
Arnold
et
al.
(
1985)

Rat
2
generations
NOAEL
 
2
mg/
kg/
day
(
reproductive
effects)
Arnold
et
al.
(
1985)

Rat
2
generations
LOAEL
 
0.016
mg/
kg/
day
(
developmental
effects,
lymphocytosis
and
fibrosis
of
the
liver)
Arnold
et
al.
(
1985)

Rat
90
weeks
LOAEL
 
10
mg/
kg/
day
(
hepatocyte
hypertrophy,
bile
duct
hyperplasia)
Smith
et
al.
(
1985)

Mouse
18
months
LOAEL
 
13
mg/
kg/
day
(
hepatocyte
hypertrophy)
Smith
et
al.(
1989)

a
NR
=
Not
Reported
44
Appendix
8.
Terrestrial
Toxicity
Data
 
Miscellaneous
Sources
Species
Test
Duration/
Type
Endpoint/
Concentration
Reference
Birds
Japanese
quail
5­
day
LC50
 
568
ppm
Hill,
E.
and
M.
Camardese,
(
1986)
Bobwhite
quail
NRa
LD50
­­
575
mg/
kg
U.
S.
National
Library
of
Medicine.
(
1995)
Mallard
duck
NR
LD50
­­
1450
mg/
kg
U.
S.
National
Library
of
Medicine.
(
1995)
Coturnix
quail
Acute
(
5
days)
NOAEL
 
22,500

g/
kg
BW­
day
Hill
and
Camardese
(
1986)
Japanese
quail
Acute
(
5
days)
 
dosed
one
time
each
day
for
study
duration
LOEL
 
500
mg/
kg
per
day
(
hepatic
toxicity,
elevation
in
level
of
hepatic
porphyrins)
Carpenter
et
al.
(
1985)
Carpenter,
Williams
and
Buhler
 
1985
Herring
gulls
(
Larus
argentatus)
Egg
injection
study
LD50
 
4300

g/
kg
(
embryo
mortality)
Boersma
et
al.,
1986
Herring
gulls
(
Larus
argentatus)
Egg
injection
study
1500

g/
kg
 
significant
reduction
in
embryonic
weight
Boersma
et
al.,
1986
Double­
breasted
Cormorant
(
Phalacrocorax
auritus)
NR
95
percent
reduction
in
egg
hatchability
from
a
contaminated
colony
 
0.01
mean
ppm
in
18
eggs,
wet
weight
Weseloh
et
al.,
1983
Mammals
Rat
Chronic/
sub­
chronic
 
30
days
LOAEL
 
1
g/
kg
BWday
(
1000
mg/
kg
BWday
 
hepatoxicity,
porphyria,
possible
alteration
of
oestrus
cycles.
Alvarez,
2000
Mink
Chronic
(
47
weeks)
LOAEL
 
1
ppm
(
0.16
mg/
kg
BW­
day)
 
reduced
birthweight
and
increased
mortality
Bleavins
et
al.,
1984
Rat
Chronic
(>
247
days)
NOAEL
 
1,600

g/
kg
BW­
day
(
reproductive
effects)
Grant
et
al.,
1977
a
NR
=
Not
Reported
45
APPENDIX
9:
Terrestrial
Exposure
Assessment
The
meadow
vole
(
Microtus
pennsylvanicus)
was
selected
as
the
mammalian
receptor
species
for
evaluating
potential
effects
of
HCB
to
mammals.
The
meadow
vole
is
primarily
herbivorous
and
is
widely
distributed
in
the
United
States
(
U.
S.
EPA,
1993b).
The
northern
bobwhite
quail
(
Colinus
virginianus)
was
selected
as
the
avian
receptor
species
for
evaluating
potential
effects
of
the
components
of
HCB
to
avian
species.
The
northern
bobwhite
quail
feeds
mainly
on
seeds
and
lowlying
vegetation.
The
bobwhite
range
includes
the
eastern
and
central
U.
S.
as
well
as
portions
of
the
Rocky
Mountains
and
the
southwest.
The
mouse
and
bobwhite
were
selected
because
a
relatively
large
proportion
of
the
diet
of
both
species
is
comprised
of
vegetation
and
there
is
an
extensive
amount
of
toxicity
data
available
for
these
species,
particularly
for
the
bobwhite.

The
following
discussion
presents
the
methods
used
to
calculate
the
potential
ingestion
of
chemicals
by
the
mouse
and
bobwhite
via
the
ingestion
of
food
(
i.
e.,
terrestrial
plants)
and
surface
soil.
The
equations
presented
below
were
derived
based
on
equations
presented
by
U.
S.
EPA
(
1989).
The
following
equation
was
used
to
calculate
the
dose
of
chemicals
that
a
mouse
or
bobwhite
would
be
expected
to
obtain
from
the
ingestion
of
terrestrial
plants:

(
1)
Dose
FI
*
C
plant
diet
=

where:

Dose
plant
=
amount
of
chemical
ingested
per
day
via
ingestion
of
plants
(
mg/
kg
bw­
d);
FI
=
food
ingestion
rate
(
kg/
kg
bw­
d);
and
C
diet
=
estimated
chemical
concentration
in
diet
(
mg/
kg).

Food
ingestion
rates
(
FI)
of
0.35
kg/
kg
bw­
d
for
adult
voles
(
U.
S.
EPA,
1993)
and
0.093
kg/
kg
bw­
d
for
adult
bobwhites
(
U.
S.
EPA,
1993)
were
used
in
the
assessment.

The
estimated
dietary
concentration
(
C
diet)
was
calculated
using
the
following
equation:

(
2)
C
P*
C
diet
p
p
=

where:

P
p
=
proportion
of
diet
consisting
of
vegetation
(
unitless);
and
C
p
=
estimated
concentration
of
contaminant
of
concern
in
vegetation
(
mg/
kg).

The
proportion
of
the
diet
(
P
p)
consisting
of
vegetation
for
both
the
vole
and
bobwhite
was
assumed
to
be
100%.
This
value
is
fairly
consistent
with
the
data
compiled
in
U.
S.
EPA
(
1993)
for
these
species.
For
both
voles
and
bobwhites
it
was
also
assumed
that
100%
of
the
plants
ingested
are
from
the
areas
with
elevated
concentrations
of
HCB.
This
assumption
is
conservative
and
may
lead
to
an
overestimate
of
potential
risks
because
the
species
are
likely
to
also
forage
in
areas
that
may
not
have
elevated
soil
concentrations.
46
The
concentration
of
a
chemical
in
a
terrestrial
plant
(
C
p)
as
fresh
weight
was
determined
using
the
following
equation:

(
3)
C
C
*
BCF
p
soil
=

where:

C
soil
=
concentration
ofcontaminant
of
concern
detected
in
surface
soil
(
mg/
kg);
and
BCF
=
bioconcentration
factor
for
the
chemical
in
terrestrial
plants
(
unitless).

The
predicted
EEC
for
HCB
in
soil
(
U.
S.
EPA,
2002)
was
used
as
the
C
soil
in
the
model
for
voles
and
bobwhites.

Bioconcentration
Factors
(
BCFs)
compiled
by
U.
S.
EPA
(
1999)
were
used
to
calculate
the
chemical
concentrations
in
plants
for
the
assessment.
Employing
equations
1,
2,
and
3,
the
estimated
dose
that
voles
and
bobwhites
would
receive
from
ingestion
of
plants
in
the
vicinity
of
pentachlorophenoltreated
utility
poles
is
provided
in
Tables
A1
and
A2.

In
addition
to
the
ingestion
of
chemicals
accumulated
in
vegetation,
bobwhites
and
voles
also
may
be
exposed
to
chemicals
through
the
inadvertent
ingestion
of
surface
soil
while
foraging
or
grooming.
The
following
equation
was
used
to
calculate
the
dose
of
chemical
that
voles
and
bobwhites
would
be
expected
to
obtain
from
the
ingestion
of
surface
soil:

(
4)
Dose
=
SI
*
C
soil
soil
where:

Dose
soil
=
amount
of
chemical
ingested
per
day
from
soil
(
mg/
kg
bw­
d);
SI
=
soil
ingestion
rate
(
kg/
kg
bw­
d);
and
C
soil
=
chemical
concentration
in
surface
soil
(
mg/
kg).

Based
on
percent
dietary
soil
ingestion
values
presented
by
Beyer
et
al.
(
1994),
it
was
assumed
that
10%
of
the
total
mass
of
both
the
vole
and
the
bobwhites'
diet
consists
of
soil.
The
percent
soil
ingestion
was
multiplied
by
the
food
ingestion
rates
(
FI)
presented
earlier
for
these
species
to
estimate
soil
ingestion
rates
(
0.035
kg/
kg
bw­
d
for
voles
and
0.0093
kg/
kg
bw­
d
for
bobwhites).
Employing
equation
4,
the
estimated
dose
voles
and
bobwhites
would
receive
from
the
ingestion
of
soil
for
HCB
are
discussed
in
the
report.

The
total
dietary
exposure
levels
for
voles
and
bobwhites
was
determined
using
the
following
equation:

(
5)
Dose
Dose
Dose
total
plant
soil
=
+
47
Using
equation
5,
the
estimated
total
dose
voles
and
bobwhites
would
be
expected
to
receive
from
the
ingestion
of
plants
and
soil
is
shown
in
Tables
A1
and
A2.

In
the
Risk
Characterization
section,
the
total
dietary
intakes
are
compared
to
toxicity
reference
values
(
e.
g.,
LD50,
NOAEL)
to
determine
if
adverse
effects
are
likely
to
occur
in
voles
and
bobwhites
from
the
ingestion
of
HCB
in
terrestrial
plants
and
surface
soil.
The
toxicity
reference
values
for
avian
species
identified
for
HCB
were
based
on
laboratory
toxicity
data
on
several
avain
species
including
the
bobwhite
quail
and
the
Japanese
quail.
Toxicity
reference
values
from
studies
based
on
avian
species
other
that
the
bobwhite
quail
were
not
adjusted
for
the
bobwhite's
body
weight
based
on
recommendations
in
Sample
et
al.
(
1996).
Toxicity
reference
values
for
mammals
were
adjusted
to
the
vole's
body
weight
using
the
following
equation
provided
in
Sample
et
al.
(
1996):

(
6)
NOAEL
NOAEL
*
(
bw
/
bw
)
w
t
tw
1/
4
=

Where:

NOAEL
w
=
NOAEL
for
mammalian
wildlife
species
(
e.
g.,
meadow
vole)
NOAEL
t
=
NOAEL
for
laboratory
mammalian
species
(
e.
g.,
rat,
rabbit,
etc.)
bw
w
=
body
weight
for
mammalian
wildlife
species
(
e.
g.,
meadow
vole)
bw
t
=
body
weight
for
laboratory
mammalian
species
(
e.
g.,
rat,
rabbit,
etc.)

The
body
weight
for
the
meadow
vole
was
assumed
to
be
0.044kg
and
body
weights
for
the
rat,
rabbit,
mouse,
and
mink
were
assumed
to
be
0.35kg,
3.8kg,
0.035kg,
and
1.0kg,
respectively,
based
on
reference
values
provided
in
Sample
et
al.
(
1996)
and
U.
S.
EPA's
Wildlife
Exposure
Factors
Handbook
(
U.
S.
EPA,
1993).

Following
adjustment
of
the
mammalian
toxicity
reference
values,
the
total
dose
for
the
bobwhite
and
vole
of
HCB
were
compared
with
the
toxicity
reference
values
for
these
contaminants
to
determine
the
risk
quotients
(
RQs).
The
RQs
for
the
bobwhite
and
vole
based
on
exposure
to
HCB
are
shown
in
Tables
A1
and
A2,
respectively.
48
Table
A1.
Acute
and
Chronic
Risk
Quotient
Calculations
for
Birds
EEC
(
mg/
kg)
Bobwhite
Soil
Ingestion
Rate
(
kg/
kg­
d)
Bobwhite
Soil
Ingestion
(
mg/
kg­
day)
BCF
(
a)
Vegetation
Residue
mg/
kg)
Proportion
of
Vegetation
in
Bobwhite's
Diet
(
unitless)
Estimated
Chemical
Concentration
in
the
Diet
(
mg/
kg)
Food
Ingestion
Rate
(
kg/
kg
bw­
d)
Dose
from
Plant
Ingestion
(
mg/
kg­
day)
Total
Dose
(
Plant
+
Soil)

(
mg/
kg­
day)
Endpoint
RQ
HCB
2.58e­
08
0.0093
2.40e­
10
2.20e­
03
5.68e­
11
1
5.68e­
11
0.093
5.28e­
12
2.45e­
10
LD50
568
(
mg/
Kg)
acute
4.32E­
13
NOAEL
1
(
mg/
kg)
chronic
2.45E­
10
(
a)
Values
for
Bioconcentration
Factors
were
provided
in
U.
S.
EPA
1999.

Table
A2.
Acute
and
Chronic
Risk
Quotient
Calculations
for
Mammals
EEC
(
mg/
kg)
Vole
Soil
Ingestion
Rate
(
kg/
kg­
d)
Vole
Soil
Ingestion
(
mg/
kg­
day)
BCF
(
a)
Vegetation
Concentrat
ion
(
mg/
kg)
Proportion
of
Vegetation
in
Vole's
Diet
(
unitless)
Estimated
Chemical
Concentration
in
the
Diet
(
mg/
kg)
Food
Ingestion
Rate
(
kg/
kg
bw­
d)
Dose
from
Plant
Ingestion
(
mg/
kg­
day)
Total
Dose
(
Plant
+
Soil)

(
mg/
kg­
day)
LD50
(
mg/
kg)
RQ
HCB
2.58e­
08
0.035
9.03e­
10
4.91e­
03
1.27e­
10
1
1.27e­
10
0.35
4.43e­
11
9.47e­
10
LD50
708
mg/
kg
acute
1.34E­
12
NOAEL
0.95
mg.
kg­
day
Chronic
9.94E­
10
(
a)
Values
for
Bioconcentration
Factors
were
provided
in
U.
S.
EPA
1999.