Document ID: EPA-HQ-OPP-2004-0402-0036
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2005-03-18T05:00Z

Page
1
of
65
ECOLOGICAL
RISK
ASSESSMENT
Dioxins
And
Furans:
Hazard
and
Risk
to
Wildlife
and
Aquatic
Organisms
3/
4/
05
Page
2
of
65
I.
Executive
Summary
Chlorinated
dibenzo­
dioxins
(
CDDs)
and
chlorinated
dibenzo­
furans
(
CDFs)
occur
in
pentachlorophenol
as
contaminants.
These
compounds
are
inherently
toxic,
as
well
as
environmentally
persistent,
and
their
presence
may
increase
the
ecological
risk
associated
with
the
use
of
pentachlorophenol.
There
are
many
congeners
of
CDDs/
CDFs,
ranging
from
monochlorinated
to
octachlorinated.
The
most
toxic
for
each
compound
seems
to
be
the
2,
3,
7,
8­
tetrachlorinated
congener,
referred
to
as
TCDD
or
TCDF
for
dioxin
or
furan,
respectively.

Data
on
the
ecological
effects
of
CDDs
and
CDFs
are
relatively
limited.
There
are
no
Guideline
studies
required
for
either
compound,
since
they
are
not
registered
as
pesticides.
Most
research
efforts
have
been
focused
primarily
on
2,3,7,8­
chlorinated
CDD/
CDFs,
especially
2,3,7,8­
TCDD.
The
majority
of
the
data
presented
in
this
chapter
are
based
on
compilations
of
data
in
U.
S.
EPA
(
1990),
U.
S.
EPA
(
1993),
and
the
U.
S.
EPA
ACQUIRE
database.

A.
Toxicity
to
Aquatic
Organisms:

In
aquatic
toxicity
studies,
the
most
toxic
congener
of
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans
is
2,3,7,8­
TCDD,
probably
due
to
its
tendency
to
be
taken
up
more
readily
than
other
congeners.
Fish
show
greater
sensitivity
to
dioxins
than
do
aquatic
invertebrates,
but
dioxins
are
very
highly
toxic
to
all
aquatic
organisms.
The
available
96­
hour
LC
50
values
for
fish
range
from
1.83
ng/
L
(
ppt)
in
rainbow
trout
(
Bol
et
al,
1989)
to
5.60
ng/
L
in
coho
salmon
(
Miller
et
al,
1973).
The
only
available
48­
hour
LC
50
value
for
aquatic
invertebrates
was
>
1,030
ng/
L
in
daphnids
(
Adams
et
al,
1986).
Longer­
term
LC
50
values
indicate
greater
toxicity
[
0.0011
ng/
L
in
rainbow
trout
(
Mehrle
et
al
,
1988)
to
1.7
ng/
L
in
fathead
minnows
(
Adams
et
al,
1986)
for
28­
day
exposures],
due
to
TCDD's
tendency
to
cause
delayed
mortality.
TCDD­
induced
mortality
typically
occurs
between
30
to
80
days
after
the
initial
exposure,
even
from
short­
term
exposure
durations.
Early
life­
stage
testing
in
fish
resulted
in
a
NOEC
value
of
0.0011
ng/
L
(
Mehrle
et
al,
1988).
Based
on
the
available
data,
the
most
sensitive
fish
life
stage
is
during
yolk
sac
resorption
in
the
developing
embryo.

Data
on
the
toxicity
of
furans
to
aquatic
organisms
indicate
that
they
are
generally
less
toxic
than
dioxins,
but
2,3,7,8­
TCDF
is
still
very
highly
toxic
to
fish
[
13­
day
LC50
value
of
16
ng/
L
in
medaka
(
Wisk
&
Cooper,
1990a)].

Many
sublethal
effects
have
also
been
reported
for
polychlorinated
dioxins
and
furans,
including:
reduced
growth;
weight
loss;
abnormal
hatching;
cellular
alterations;
numerous
types
of
lesions
in
a
broad
spectra
of
organs;
and
behavioral
effects
on
swimming,
reduced
feeding,
and
loss
of
response
to
external
stimuli.

B.
Toxicity
to
Terrestrial
Organisms:
Page
3
of
65
Avian
toxicity
data
are
more
limited
for
the
CDDs/
CDFs.
The
available
LD
50
values
for
2,3,7,8­
TCDD
range
from
15

g/
kg
in
Northern
bobwhite
to
810

g/
kg
in
Ringed
turtle­
dove
(
Hudson
et
al,
1984),
all
of
which
indicate
very
high
acute
toxicity.
An
LC
50
of
167
ppt
for
Northern
bobwhite
in
an
8­
day
dietary
exposure
study
also
demonstrates
the
very
high
toxicity
of
TCDD
to
birds.
Studies
examining
effects
of
injecting
TCDD
into
eggs
have
shown
that
very
low
levels
(
NOEC
of
0.1

g/
kg)
cause
mortality
in
chicks
up
to
28­
days
post­
hatch
(
Nosek
et
al,
1992c).

An
LC
50
of
0.264

g/
kg
bw/
day
in
female
mink
(
Hochstein
et
al,
1988)
indicates
that
TCDD
has
very
high
acute
toxicity
to
mammals.
Studies
of
the
reproductive
effects
of
TCDD
in
monkeys
have
shown
it
to
cause
spontaneous
abortions
and
other
toxic
effects
at
very
low
levels
(
Kociba
and
Schwetz,
1982).

CDDs
and
CDFs
are
very
highly
toxic
to
birds,
mammals
and
aquatic
organisms.
CDDs
(
and
possibly
furans)
are
capable
of
producing
lasting
toxic
effects;
even
a
relatively
short
exposure
to
TCDD
(
as
little
as
6
hours)
can
result
in
mortality
occurring
as
much
as
80
days
later.
TCDD
is
a
known
endocrine
disruptor,
and
it
is
likely
that
other
dioxin
congeners
and
furans
produce
similar
effects.
While
there
are
no
EPA
Guideline
studies
available
for
these
compounds,
the
literature
data
are
sufficient
to
conclude
that
these
compounds
have
the
potential
to
adversely
affect
wildlife
and
aquatic
organisms.

C.
Ecological
Risk
There
is
a
great
deal
of
uncertainty
in
these
quantitative
estimates,
however,
due
to
a
lack
of
standard
toxicity
endpoints
and
exposure
information.
There
are
also
uncertainties
and
limitations
inherent
to
the
exposure
models
used,
as
described
in
the
Environmental
Modeling
chapter
of
this
document.
Based
on
the
information
that
is
available,
it
is
unlikely
that
the
CDDs/
CDFs
resulting
from
the
use
of
pentachorophenol
on
utility
poles
are
posing
an
immediate
risk
to
wildlife
or
aquatic
organisms;
however,
it
is
possible
that
the
buildup
of
these
persistent
and
bioaccumulative
compounds
over
time
may
eventually
reach
levels
that
pose
a
risk
to
these
organisms.
Additionally,
pentachlorophenol
is
only
one
of
many
sources
of
CDDs/
CDFs
in
the
environment;
it
is
not
currently
possible
to
quantify
the
portion
of
the
aggregate
environmental
risk
from
CDDs/
CDFs
that
is
attributable
to
the
use
of
pentachlorophenol
on
utility
poles.

II.
Background/
Introduction
CDD/
CDFs
are
dispersed
throughout
the
global
ecosystem.
They
are
chemically
stable
and
can
bioaccumulate
in
animal
tissue.
Data
on
the
ecological
effects
of
CDD/
CDFs
are
limited.
Most
research
efforts
have
been
focused
primarily
on
2,3,7,8­
chlorinated
CDD/
CDFs,
especially
2,3,7,8­
TCDD.
This
chapter
presents
a
compilation
and
evaluation
of
the
available
ecological
effects
data
for
CDD/
CDFs
on
aquatic
life
(
e.
g.,
fish,
arthropods,
etc.),
wild
mammals
(
e.
g.,
mink,
monkeys),
and
birds.

The
majority
of
the
data
presented
in
this
chapter
are
based
on
compilations
of
data
in
U.
S.
EPA
(
1990),
U.
S.
EPA
(
1993),
and
the
U.
S.
EPA
ACQUIRE
database.
In
some
instances,
data
from
the
same
study
are
presented
in
multiple
tables.
A
brief
discussion
of
some
selected
studies
is
also
Page
4
of
65
presented.
The
reader
is
referred
to
U.
S.
EPA
(
1990)
and
U.
S.
EPA
(
1993)
for
a
more
detailed
presentation
of
the
studies
used
therein.

III.
Toxicity
of
CDDs/
CDFs
to
Aquatic
Organisms
While
there
are
no
EPA
Guideline
studies
available
for
CDDs/
CDFs,
literature
data
provide
some
studies
with
comparable
endpoints.
These
values
are
presented
in
Table
1,
below.

Table
1:
Selected
Aquatic
Toxicity
Values
for
CDDs/
CDFs
Compound
Species
Test
Type/
Duration
Endpoint
Toxicity
classification
Citation
2,3,7,8­
TCDD
Coho
salmon
96­
hour
acute
LC50=
5.60
ppt
very
highly
toxic
Miller
et
al,
1973
Rainbow
trout­­
sac
fry
96­
hour
acute
LC50=
1.83
ppt
very
highly
toxic
Bol
et
al,
1989
Rainbow
trout­­
swimup
fry
28­
day
early
life­
stage
LC50=
0.046
ppt
NOEC=
0.0011
ppt
very
highly
toxic
Mehrle
et
al,
1988
fathead
minnows
28­
day
LC50=
1.7
ppt
very
highly
toxic
Adams
et
al,
1986
Daphnia
magna
48­
hour
acute
EC50>
1020
ppt
very
highly
toxic
Adams
et
al,
1986
Annelid
worm
55
days
NOEC
=
200
ppt
Miller
et
al,
1973
Snail
12
days
NOEC
=
200
ppt
Miller
et
al,
1973
Dibenzofuran
Sheepshead
minnow
96­
hour
acute
LC50=
1800
ppb
moderately
toxic
Heitmuller
et
al,
1981
fathead
minow
96­
hour
acute
LC50=
1140
ppb
moderately
toxic
Brooke,
1991
Daphnia
magna
48­
hour
acute
LC50=
1340
ppb
moderately
toxic
Maas,
1990
Mysid
96­
hour
acute
LC50=
1310
ppb
moderately
toxic
EPA,
1978
Diatom,
Skeletonema
costatum
96­
hour
acute
EC50=
1500
ppb
EPA,
1978
2,3,7,8­
TDCF
Medaka
13­
day
LC50=
16
ppb
very
highly
toxic
Wisk
&
Cooper,
1990a
Based
on
the
available
data,
the
most
toxic
CDD
congener
is
2,3,7,8­
TCDD
and
the
most
toxic
CDF
congener
is
2,3,7,8­
TCDF.
Many
sublethal
effects
have
also
been
reported
for
CDDs
and
Page
5
of
65
CDFs,
including:
reduced
growth;
weight
loss;
abnormal
hatching;
cellular
alternations;
numerous
types
of
lesions
in
a
broad
spectra
of
organs;
and
behavioral
effects
on
swimming,
reduced
feeding,
and
loss
of
response
to
external
stimuli.

A.
Aquatic
Toxicity
of
CDDs
The
available
studies
on
CDDs
and
CDFs
published
between
1984
and
1989
were
collected
and
reviewed
in
a
document
entitled
"
Background
Document
to
the
Integrated
Risk
Assessment
for
Dioxins
and
Furans
from
Chlorine
Bleaching
in
Pulp
and
Paper
Mills"
(
U.
S.
EPA,
1990).
This
1990
EPA
report
evaluated
the
aquatic
toxicity
literature
published
subsequent
to
the
comprehensive
EPA
reviews
in
the
ambient
water
quality
criteria
document
(
U.
S.
EPA,
1984a)
and
the
health
and
environmental
effects
profile
(
U.
S.
EPA,
1984b).
Thus,
the
information
contained
in
U.
S.
EPA
(
1990)
does
not
include
a
detailed
review
of
earlier
aquatic
toxicity
data,
but
does
contain
a
minimal
amount
of
data
from
studies
prior
to
this
timeframe.
Test
concentrations
and
effects
data
reported
in
U.
S.
EPA
(
1990)
are
summarized
for
CDDs
in
Appendix
1.

The
document,
"
Interim
Report
on
Data
and
Methods
for
Assessment
of
2,3,7,8­
Tetrachlorodibenzo­
p­
dioxins
Risks
to
Aquatic
Life
and
Associated
Wildlife"
(
U.
S.
EPA,
1993)
provides
a
compilation
of
data
on
the
toxic
effects
of
2,3,7,8­
TCDD
to
aquatic
life.
A
summary
of
this
compilation
is
provided
in
Appendix
2.
This
report
was
prepared
as
part
of
the
Agency's
effort
to
conduct
a
scientific
reassessment
of
the
risk
of
TCDD.

U.
S.
EPA
(
1993)
conducted
a
literature
search
to
obtain
aquatic
toxicity
and
accumulation
data
for
both
fresh
and
saltwater
organisms.
Laboratory
tests
and
aquatic
model
ecosystem
studies
were
included.
Most
of
the
studies
involved
short­
term
exposure
periods
where
the
organisms
were
exposed
to
TCDD.
Appendix
2
contains
the
available
data
for
freshwater
plants,
freshwater
fish,
freshwater
invertebrates,
amphibians,
and
saltwater
fish.

EPA
maintains
a
database
(
ACQUIRE)
where
ecological
data
from
laboratory
studies
on
various
chemicals
are
stored.
Summary
data
for
aquatic
toxicity
from
the
EPA
ACQUIRE
database
are
presented
in
Appendix
3.
The
data
presented
are
only
those
data
where
a
toxic
endpoint
was
provided.

The
most
toxic
CDD
to
aquatic
life
is
2,3,7,8­
TCDD.
CDDs
appear
to
be
more
toxic
to
fish
than
to
invertebrates;
however,
laboratory
data
indicate
that
TCDD
is
very
highly
toxic
to
all
aquatic
organisms.

The
mode
of
action
of
2,3,7,8­
TCDD
is
unknown,
but
it
appears
to
be
related
to
neural
toxicity
and
may
also
affect
the
immune
system,
like
many
other
organochlorine
chemicals.
The
pattern
of
TCDD­
induced
mortality
typically
occurs
between
30
to
80
days
after
the
initial
exposure,
even
for
a
duration
of
exposure
as
short
as
6
hours.
Mortality
in
fish
appears
to
be
a
function
of
both
exposure
duration
and
the
test
concentration.
The
available
bioconcentration
tests
were
not
of
sufficient
duration
to
achieve
a
steady­
state
bioconcentration
factor
(
BCF).
All
BCF
values
have
been
estimated
using
the
uptake
rate
(
k
1)
and
depuration
rate
(
k
2)
values.
Page
6
of
65
Aquatic
studies
with
2,3,7,8­
TCDD
have
generally
been
conducted
either
as
continuous
long­
term
exposures
or
short­
term
exposures
(
6
to
96
hours)
with
long­
term
observation
periods.
The
long
observation
periods
are
necessary
to
monitor
the
delayed
mortality
(
e.
g.,
30
to
80
days
postexposure
which
is
characteristic
of
2,3,7,8­
TCDD.
The
occurrence
of
delayed
effects
has
some
serious
implications
for
both
hazard
evaluations
and
risk
assessments.
For
example,
the
reported
results
of
typical
acute
toxicity
tests
are
probably
not
an
accurate
indication
of
toxicity
because
of
the
short
observation
periods.
In
risk
scenarios,
sporadic
acute
releases
may
cause
serious
delayed
effects.
Of
most
concern
is
that
there
are
insufficient
data
to
assess
the
effects
from
the
many
complex
combinations
of
exposure
concentrations
and
the
various
durations
of
the
exposure
that
occur
in
nature.

The
delayed
mortality
demonstrated
in
dioxin
studies
indicates
that
the
most
appropriate
method
of
reporting
the
effects
of
dioxin
is
not
a
simple
96­
hour
LC
50
value
or
a
maximum
acceptable
toxicant
concentration
(
MATC)
range.
The
toxicity
of
2,3,7,8­
TCDD
is
a
function
of
the
duration
of
the
exposure,
as
well
as
the
dose
level.
At
any
given
test
concentration,
continuous
exposures
appear
to
produce
toxic
effects
in
a
shorter
time
than
do
short­
term
exposures.
Dioxin
effects
from
various
routes
of
exposure,
including
water,
sediment,
diet,
and
intraperitoneal
injection
are
summarized
in
Table
1.
Based
on
these
data,
exposure
via
water
appears
to
be
the
most
toxic,
single
route
of
exposure.
However,
it
is
likely
that
the
degree
of
an
effect
could
be
increased
by
a
combination
of
exposure
routes,
such
as
would
be
found
in
a
natural
ecosystem.
This
is
especially
likely
since
the
low
water
solubility
and
high
octanol/
water
partition
coefficient
of
2,3,7,8­
TCDD
suggest
most
of
that
which
would
be
found
in
the
aquatic
ecosystem
would
be
sorbed
to
or
bioaccumulated
in
dietary
sources.

The
duration
of
an
exposure
is
important
when
evaluating
the
inherent
chronic
toxicity
of
2,3,7,8­
TCDD;
however,
toxic
effects
were
reported
even
for
the
shortest
2,3,7,8­
TCDD
exposure
period
in
water
(
6
hours),
following
which
rainbow
trout
mortalities
occurred
between
78
and
139
days
after
the
exposure
(
Branson
et
al.,
1985).
Adams
et
al.
(
1986)
estimated
that
90
percent
of
steady­
state
uptake
of
2,3,7,8­
TCDD
in
rainbow
trout
would
take
longer
than
48
days.
Consequently,
an
adequate
chronic
study
would
require
than
an
exposure
period
be
extended
until
the
maximum
level
of
2,3,7,8­
TCDD
has
been
reached,
plus
sufficient
additional
time
for
effects
to
be
produced.
Review
of
the
studies
in
Table
1
indicates
that
the
studies
with
water
exposures
are
not
of
sufficient
duration
for
a
steady­
state
concentration
of
2,3,7,8­
TCDD
to
have
been
achieved
in
fish
tissues.

The
longest
exposure
to
2,3,7,8­
TCDD
via
water
was
reported
to
be
28
days,
with
an
additional
28­
days
post­
treatment
for
observation
for
rainbow
trout
fry
(
Mehrle
et
al.,
1988).
In
that
study,
a
NOEC
was
not
determined.
At
the
lowest
test
concentration
(
0.038
ng/
l),
45
percent
mortality
occurred
before
the
test
ended.

The
28­
day
trout
study
(
Mehrle
et
al.,
1988)
did
not
investigate
deposition
of
2,3,7,8­
TCDD
in
the
eggs
by
the
female.
Some
evidence
is
available
which
suggests
that
for
trout
the
most
sensitive
life
stage
for
chlorinated
chemicals
occurs
when
the
toxicant
stored
in
the
yolk
is
sorbed
into
the
developing
embryo
at
the
end
of
the
yolk
sac
stage.
Similarly,
Helder
(
1980)
reported
that
following
a
96­
hour
exposure
to
2,3,7,8­
TCDD,
the
highest
mortality
in
freshly
fertilized
pike
eggs
occurred
during
resorption
of
the
yolk.
Mortality
reached
almost
100
percent
at
2,3,7,8­
Page
7
of
65
TCDD
exposure
concentration
of
10
ng/
l.
The
NOEC
was
reported
as
less
than
the
lowest
test
concentration
(
0.1
ng/
l).

Sublethal
effects
of
CDDs
have
been
studied
by
several
researchers,
but
it
appears
that
the
studies
were
not
conducted
for
sufficient
duration
to
determine
the
inherent
toxicity
of
CDDs.
The
effects
of
2,3,7,8­
TCDD
on
the
development
of
medaka
embryos
have
been
studied
by
Wisk
and
Cooper
(
1986).
It
was
found
that
20
to
40

g/
l
caused
tube
hearts,
hemostasis,
and
liver
necrosis.
In
another
study
(
Helder
and
Seinen,
1986),
2,3,7,8­
TCDD
and
2,3,7,8­
TCDF
produced
lethal
edematous
conditions
in
rainbow
trout.
Similar
effects
of
2,3,7,8­
TCDD
were
reported
by
Helder
(
1980)
during
early
life
stages
of
the
pike,
in
which
the
sublethal
effects
included:
reduced
size
at
hatch,
tail­
first­
hatching,
alteration
of
blood
vessel
walls,
pericardial
edema,
enlarged
nuclei
and
degradation
in
hepatocytes,
and
severe
generalized
edemas
preceding
death.

Sublethal
effects
on
rainbow
trout
have
been
studied
by
Spitsberg
et
al.
(
1988a)
following
single
intraperitoneal
injections
of
either
1,
5,
25,
or
125

g
2,3,7,8­
TCDD/
kg.
At
both
25
and
125

g/
kg,
mortality
occurred
before
body
weight
loss
could
be
detected.
At
5

g/
kg,
effects
included
reductions
in
activity,
feeding,
and
growth.
Moderate
to
marked
leukopenia
and
thrombocytopenia
were
also
found.
The
2,3,7,8­
TCDD­
treated
trout
also
suffered
higher
stress
from
handling
than
controls.
Numerous
types
of
lesions
were
reported
in
lymphomyeliod
tissues
of
the
thymus,
spleen,
and
kidneys,
as
well
as
epithelial
lesions
of
the
stomach
mucosa,
liver,
pancreas,
gill,
and
skin.
The
types
of
lesions
included
the
following:
thymic
involution;
splenic
lymphoid
depletion;
decreased
kidney
hematopoiesis;
multifocal
necrosis,
atrophy,
and
hyperplasia
of
the
stomach
mucosa;
mild
to
severe
hepatocyte
vacuolation
and
ballooning
degeneration
of
the
liver
and
hyperplasia
of
the
bile­
duct;
vacuolation
of
pancreatic
exocrine
cells;
mild
fusion
of
gill
lamellae;
and
necrosis
of
fin
margins.
Histologic
changes
in
epithelial
organs
resembled
the
lesions
found
in
early
stage
studies
of
rainbow
trout
exposed
to
water­
borne
2,3,7,8­
TCDD.

Sublethal
effects
in
yellow
perch
following
single
intraperitoneal
injections
of
either
1,
5,
25,
or
125

g
2,3,7,8­
TCDD/
kg
have
also
been
reported
(
Spitsberg
et
al.,
1988b).
Many
effects
were,
in
general,
similar
to
those
reported
in
rainbow
trout.
Yellow
perch
did
not,
however,
show
lethality
resulting
from
handling
as
seen
in
rainbow
trout.
Lesions
in
the
thymus,
spleen,
kidney,
stomach,
and
skin
are
similar
to
the
effects
observed
in
rainbow
trout.
Cardiac
lesions
not
seen
in
rainbow
trout
were
found,
including:
necrosis
of
myocytes
subjacent
to
the
epicardial
surface
of
the
ventricle;
fibrinous
pericarditis;
and
hypertrophy
and
hyperplasis
of
the
pericardial
mesothelium.
Many
of
the
histologic
lesions,
including
fibrinous
pericarditis,
have
also
been
found
in
mammals,
chickens,
or
turkeys
treated
with
2,3,7,8­
TCDD.

Spitsberg
et
al.
(
1986)
measured
immune
responses
in
rainbow
trout
at
14
and
30
days
after
single
intraperitoneal
injections
of
either
0.1,
1.0,
or
10

g/
kg
of
2,3,7,8­
TCDD.
Trout
injected
with
2,3,7,8­
TCDD
at
either
0.1
or
1.0

g/
kg
remained
clinically
normal.
Trout
treated
with
10

g/
kg
of
2,3,7,8­
TCDD
became
hypophagic
and
exhibited
fin
necrosis,
ascites,
and
suppression
of
hematopoiesis.
Concanavalin
A­
induced
blastogenesis
of
thymic
and
splenic
lymphocytes
were
not
significantly
changed,
however,
suppression
of
the
pokeweed
mitogen­
induced
response
of
splenic
lymphocytes
occurred.
No
statistically
significant
alterations
occurred
in
humoral
immune
responses,
and
phagocytic
activity
of
peritoneal
macrophages
was
not
decreased.
In
rainbow
Page
8
of
65
trout,
immunosupression
was
evident
only
at
2,3,7,8­
TCDD
doses
approaching
20

g/
kg,
which
is
the
80­
day,
single­
dose,
parenteral
LD
50
value.

Spitsberg
et
al.
(
1988c)
found
no
statistically
significant
effects
on
rainbow
trout
mortality
or
mean
time­
until­
death
following
a
challenge
with
infectious
haematopoietic
necrosis
virus
(
IHNV).
Rainbow
trout
were
injected
intraperitoneally
with
single
doses
of
either
0.01,
0.1,
1.0,
or
10

g/
kg
2,3,7,8­
TCDD.
In
virus­
free
trout,
TCDD­
induced
effects
first
appeared
4
to
5
weeks
after
injection
of
10

g/
kg.
The
effects
included
fin
necrosis,
as
well
as
reduction
in
activity
and
food
consumption.
No
deaths
occurred
in
the
virus­
free
rainbow
trout
during
the
5
weeks
after
treatment.

Behavioral
effects
induced
by
2,3,7,8­
TCDD
in
rainbow
trout
have
also
been
reported
by
Mehrle
et
al.
(
1988).
Rainbow
trout
were
severely
stressed
at
all
test
concentrations
(
i.
e.,
0.038
to
0.789
ng/
l
of
2,3,7,8­
TCDD).
Behavioral
impairments
increased
over
time
and
with
increasing
test
concentration.
The
behavioral
changes
included:
lethargic
swimming,
abnormal
head­
up
swimming
posture,
feeding
inhibition,
and
lack
of
response
to
external
stimuli.
All
of
these
behaviors
would
increase
the
likelihood
of
predation
in
a
natural
ecosystem.

B.
Aquatic
Toxicity
of
CDFs
Data
compiled
in
U.
S.
EPA
(
1990)
on
the
toxic
effects
of
CDFs
are
summarized
in
Appendix
4,
and
a
summary
of
aquatic
toxicity
data
for
CDFs
from
the
ACQUIRE
database
is
presented
in
Appendix
5..
The
longest
exposure
of
2,3,7,8­
TCDF
via
water
was
28
days,
with
an
additional
28­
days
post­
treatment
for
observation
of
the
rainbow
trout
fry
(
Mehrle
et
al.,
1988).
The
NOEC
for
2,3,7,8­
TCDF
was
reported
to
be
0.41
ng/
l.
It
appears
that
2,3,7,8­
TCDF
levels
in
the
trout
may
have
reached
equilibrium
at
the
higher
concentration
(
3.93
ng/
l),
but
it
is
not
evident
that
equilibrium
was
achieved
in
28
days
at
the
lower
concentration
(
0.41
ng/
l).
Consequently,
it
is
uncertain
whether
the
only
study
available
concerning
2,3,7,8­
TCDF
was
of
sufficient
duration
to
produce
maximum
toxic
effects.
The
behavioral
effects
induced
by
2,3,7,8­
TCDF
were
similar
to
those
observed
in
2,3,7,8­
TCDD
exposures;
however,
the
observed
responses
were
of
lesser
magnitude
(
Mehrle
et
al.,
1988).

C.
Conclusions
Concerning
Aquatic
Toxicity
of
CDDs
and
CDFs
CDDs
appear
to
be
more
toxic
to
fish
than
to
aquatic
invertebrates;
however,
the
available
toxicity
values
demonstrate
that
CDDs
are
very
highly
toxic
to
all
the
aquatic
organisms
that
were
tested.
Many
aquatic
tests
have
been
conducted
with
2,3,7,8­
TCDD,
but
most
of
these
studies
were
either
bioavailability
tests
or
short­
term
exposure
studies
with
long
post­
exposure
observation
periods.
The
available
test
data
on
2,3,7,8­
TCDD
indicate
that
none
of
the
studies
are
adequate
to
define
an
acceptable,
no­
observed­
effect
concentration
(
NOEC).
Only
one
study
exposed
a
fish
to
2,3,7,8­
TCDD
in
water
for
a
reasonable
duration
(
28
days).
Even
then,
the
exposure
duration
in
the
rainbow
trout
study
was
only
about
half
of
the
48
days
estimated
to
be
required
to
achieve
90
percent
of
the
steady­
state
BCF
level
in
fish.
The
NOEC
was
less
than
0.038
ng/
l,
a
concentration
which
produced
45
percent
mortality
in
rainbow
trout
fry.

A
definitive
NOEC
has
not
been
reported
for
2,3,7,8­
TCDD.
Even
the
lowest
test
concentration
(
0.038
ng/
l)
produced
45
percent
mortality
in
rainbow
trout
exposed
to
2,3,7,8­
TCDF
for
28
Page
9
of
65
days.
The
reported
NOEC
value
for
2,3,7,8­
TCDF
is
0.41
ng/
l,
but
this
is
also
uncertain
because
of
the
limited
duration
of
observation.
The
toxicity
data
available
on
2,3,7,8­
TCDD,
and
possibly
2,3,7,8­
TCDF,
do
not
adequately
define
the
inherent
toxicity
of
these
substances
for
two
reasons:
(
1)
the
exposure
periods
are
of
insufficient
duration
for
a
steady­
state
equilibrium
to
be
reached;
and
(
2)
the
studies
do
not
address
toxic
effects
on
developing
embryos
resulting
from
deposition
of
either
2,3,7,8­
TCDD
or
2,3,7,8­
TCDF
in
the
eggs
by
the
female.

Data
on
the
toxicity
of
furans
to
aquatic
organisms
indicate
that
they
are
generally
less
toxic
than
dioxins,
but
2,3,7,8­
TCDF
is
still
very
highly
toxic
to
fish
[
13­
day
LC50
value
of
16
ng/
L
in
medaka
(
Wisk
&
Cooper,
1990a)].

The
most
sensitive
stage
in
the
life
of
a
rainbow
trout
for
chlorinated
chemicals
appears
to
occur
during
resorption
of
the
yolk
sac
in
developing
embryos.
During
this
stage,
the
lipophilic
substances
(
e.
g.,
CDDs
and
CDFs)
deposited
in
yolk
by
the
female
become
more
concentrated
in
the
yolk
on
a
per­
unit­
weight
basis
and
they
are
resorbed.
This
occurs
as
the
fry
metabolizes
and
incorporates
nutrients
contained
in
the
yolk.
As
the
toxic
substance
is
released
during
metabolism
of
the
yolk,
the
amount
of
substance
entering
the
blood
stream
reaches
a
higher
concentration
than
would
otherwise
normally
be
found.
It
is
at
this
stage
of
development
of
organisms
that
the
greatest
sensitivity
has
been
observed
for
substances
such
as
DDE
in
both
birds
and
fish.

IV.
Toxicity
Of
CDDs
and
CDFs
to
Birds
and
Terrestrial
Mammals
The
adverse
effects
to
individual
wildlife
species
from
2,3,7,8­
TCDD
have
been
documented
in
laboratory
studies.
Using
the
results
of
these
studies
to
estimate
effects
on
wild
populations
has
limitations
because
the
route
and
medium
of
administration
and
the
duration
of
exposure
to
2,3,7,8­
TCDD
for
laboratory
animals
usually
will
differ
from
that
of
wild
animals.
Using
these
studies
to
assess
effects
on
wild
species
assumes
that
the
wild
species
are
as
sensitive
or
more
sensitive
to
2,3,7,8­
TCDD
than
the
laboratory
species.
The
methodologies
for
predicting
the
effects
of
chemicals
on
terrestrial
wildlife
populations
and
ecosystems,
however,
are
still
in
development.

A.
Toxicity
Assessment
for
Birds
and
Bird
Eggs
1.
Avian
Toxicity
Data
Used
in
This
Risk
Assessment
While
there
are
no
EPA
Guideline
studies
submitted
for
CDDs/
CDFs,
the
literature
data
provide
TCDD
toxicity
values
for
some
comparable
endpoints.
These
are
provided
in
Table
2,
below.

Table
2:
Selected
Avian
Toxicity
Values
for
2,3,7,8­
TCDD
Species
Test
type/
duration
Endpoint
Toxicity
Classification
Citation
Northern
bobwhite
Acute
oral
LD50=
15

g/
kg
very
highly
toxic
Hudson
et
al,
1984
Northern
bobwhite
8­
day
dietary
LC50=
167
ppt
very
highly
toxic
Kenaga
and
Norris,
1983
Page
10
of
65
Ringneck
pheasant
10
week
reproduction
NOAEL=
0.014

g/
kg­
day
(
fertility,
embryo
mortality)
Nosek
et
al.,
1992b
and
1993
A
full
compilation
of
the
toxicity
data
for
2,3,7,8­
TCDD
in
birds
and
bird
eggs
from
U.
S.
EPA
(
1990)
is
provided
in
Appendix
6.
The
data
from
U.
S.
EPA
(
1993)
for
birds
are
presented
in
Appendix
7.

2.
Summaries
of
Additional
Avian
Toxicity
Data
2,3,7,8­
TCDD
was
administered
at
100
ng/
kg
body
weight/
day
in
a
corn
oil/
acetone
vehicle
to
3­
day
old
white
leghorn
chickens
(
Schwetz
et
al.,
1973).
This
dose
was
administered
for
21
days
and
produced
no
adverse
effects.
It
was
assumed
that
the
2,3,7,8­
TCDD
was
100
percent
absorbed
from
the
corn
oil/
acetone
vehicle.
However,
it
is
possible
that
the
absorption
of
2,3,7,8­
TCDD
from
laboratory
feed
or
food
sources
for
wild
animals
would
not
be
the
same
as
the
assumed
100
percent
absorption
of
2,3,7,8­
TCDD
from
a
corn
oil/
acetone
vehicle.

Bird
eggs
can
contain
2,3,7,8­
TCDD
transferred
from
the
mother's
body
burden
of
2,3,7,8­
TCDD.
Eggs
are
an
important
model
to
consider
in
determining
a
toxicity
endpoint
because
of
the
sensitivity
of
eggs
to
2,3,7,8­
TCDD.
Sullivan
et
al.
(
1987)
concluded
that
the
LOAEL
for
chicken
embryos
is
65
pg/
g
in
the
egg
(
65
ppt),
based
on
a
study
that
found
a
2­
fold
increase
in
cardiovascular
malformations
in
chicken
embryos
at
an
estimated
egg
concentration
of
65
pg/
g.
Although
effects
were
found
at
lower
concentrations
of
2,3,7,8­
TCDD,
the
study
concluded
that
the
evidence
for
effects
at
these
lower
levels
was
inconclusive;
thus,
the
65
ppt
value
can
be
used
in
risk
assessment
context
for
comparison
with
predicted
egg
concentrations
for
wild
species.

B.
Toxicity
Assessment
for
Terrestrial
Mammals
1.
Mammalian
Data
Used
in
This
Risk
Assessment
While
there
are
no
EPA
Guideline
studies
submitted
for
CDDs/
CDFs,
the
literature
data
provide
TCDD
toxicity
values
for
some
comparable
endpoints.
These
are
provided
in
Table
3,
below.

Table
3:
Selected
Mammalian
Toxicity
Values
for
2,3,7,8­
TCDD
Species
Test
type/
duration
Endpoint
Citation
Mouse
acute
oral
LD50=
114­
284

g/
kg
Kociba
and
Schwetz,
1982
Rat
reproduction
 
exp
osed
days
6­
15
of
gestation
NOAEL
(
litter
size
and
pup
weight)=
0.125

g/
kg­
day
Khera
and
Ruddick,
1973
A
summary
of
toxic
effects
of
2,3,7,8­
TCDD
to
wild
mammals
from
U.
S.
EPA
(
1990)
is
presented
in
Appendix
8.
Summary
data
from
U.
S.
EPA
(
1993)
are
presented
in
Appendix
9.
Page
11
of
65
2.
Summaries
of
Some
Additional
Mammalian
Data
In
a
previous
study,
Murray
et
al.
(
1979)
dosed
rats
at
100,
10,
and
1
ng/
kg/
day
through
the
diet,
and
studied
the
effects
on
subsequent
generations.
At
the
10
ng/
kg/
day
level,
Murray
et
al.
(
1979)
found
decreased
fertility
in
the
f
1
and
f
2
generations.
A
NOAEL
level
of
1
ng/
kg/
day
was
reported.

For
larger
mammals,
the
expected
dose
for
wild
species
is
compared
to
the
lowest
dose
observed
to
produce
adverse
reproductive
effects
in
rhesus
monkeys.
Schwetz
et
al.
(
1973)
reported
that
rhesus
monkeys
were
given
1.7
ng/
kg
body
weight
of
2,3,7,8­
TCDD
in
the
diet.
Of
the
seven
pregnancies
which
occurred,
four
resulted
in
chemical­
induced
abortions.

In
both
laboratory
studies,
it
was
assumed
that
absorption
from
a
laboratory
diet
is
similar
to
the
absorption
from
a
wild
diet,
and
that
these
doses
are
directly
comparable
to
the
daily
dose
to
wild
species
from
the
ingestion
of
prey
items.

V.
Ecological
Hazard
Conclusions
In
aquatic
toxicity
studies,
the
most
toxic
congener
of
polychlorinated
dibenzo­
p­
dioxins
and
dibenzofurans
appears
to
be
2,3,7,8­
TCDD,
probably
due
to
its
tendency
to
be
taken
up
more
readily
than
other
congeners.
Dioxins
appear
to
be
more
toxic
to
fish
than
to
aquatic
invertebrates,
but
are
considered
to
be
very
highly
toxic
to
all
aquatic
organisms
tested.

The
mode
of
action
of
2,3,7,8­
TCDD
is
unknown,
but
it
appears
to
be
related
to
neural
toxicity
and
may
also
affect
the
immune
system,
like
many
other
organochlorine
chemicals.
TCDDinduced
mortality
typically
occurs
between
30
to
80
days
after
the
initial
exposure,
even
from
exposure
durations
as
short
as
6
hours.
Mortality
in
fish
appears
to
be
a
function
of
both
exposure
duration
and
the
test
concentration.

Many
sublethal
effects
have
been
reported
for
polychlorinated
dioxins
and
furans.
2,3,7,8­
TCDD
elicits
a
broad
range
of
toxic
effects
which
include:
reduced
growth;
weight
loss;
abnormal
hatching;
cellular
alterations;
numerous
types
of
lesions
in
a
broad
spectra
of
organs;
and
behavioral
effects
on
swimming,
reduced
feeding,
and
loss
of
response
to
external
stimuli.

Continuous
exposures
to
fish
appear
to
produce
toxic
effects
more
rapidly
than
short­
term
exposure
to
the
same
test
concentration.
Based
on
the
available
data,
the
most
sensitive
fish
life
stage
is
during
yolk
sac
resorption
in
the
developing
embryo.

The
2,3,7,8­
TCDD
toxicological
endpoints
for
fresh
water
fishes,
marine
and
estuarine
fishes,
and
invertebrates
are
provided
in
Appendices
10,
11,
and
12,
respectively.

The
lowest
NOEL
for
2,3,7,8­
TCDD
exposure
to
birds
is
100
ng/
kg/
day
from
an
oral
dose
to
white
leghorn
chickens
over
a
21­
day
exposure
period.
For
wild
mammals,
a
NOAEL
of
1
ng/
kg/
day
in
the
rat
has
been
reported
in
a
multi­
generation
effects
study.

The
acute
and
chronic
2,3,7,8­
TCDD
toxicological
endpoints
for
birds
and
mammals
are
summarized
in
Appendices
13,
14,
15,
and
16,
respectively.
Page
12
of
65
VI.
Environmental
Risk
Assessment
Risk
characterization
integrates
the
results
of
the
exposure
and
ecotoxicity
data
to
evaluate
the
likelihood
of
adverse
ecological
effects.
The
means
of
this
integration
is
called
the
quotient
method.
Risk
quotients
(
RQs)
are
calculated
by
dividing
exposure
estimates
by
acute
and
chronic
ecotoxicity
values.

RQ
=
EXPOSURE/
TOXICITY
RQs
are
then
compared
to
OPP's
Levels
of
Concern
(
LOCs).
These
LOCs
are
used
by
OPP
to
analyze
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
may
be
mitigated
through
restricted
use
classification,
(
3)
acute
endangered
species
­
endangered
species
may
be
adversely
affected,
and
(
4)
chronic
risk
­
the
potential
for
chronic
risk
is
high
regulatory
action
may
be
warranted.
Currently,
EFED
does
not
perform
assessments
for
chronic
risk
to
plants,
acute
or
chronic
risks
to
nontarget
insects,
or
chronic
risk
from
granular/
bait
formulations
to
birds
or
mammals.

The
ecotoxicity
test
values
(
measurement
endpoints)
used
in
the
acute
and
chronic
risk
quotients
are
derived
from
required
studies.
Examples
of
ecotoxicity
values
derived
from
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
long­
term
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
generally
is
used
as
the
ecotoxicity
test
value
in
assessing
chronic
effects,
although
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
assessing
chronic
effects
to
fish
and
aquatic
invertebrates.
However,
the
NOEC
is
used
if
the
measurement
end
point
is
production
of
offspring
or
survival.

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

Table
4:
Risk
Presumptions
for
Terrestrial
Animals
Risk
Presumption
RQ
LOC
Birds
Acute
High
Risk
EEC1/
LC50
or
LD50/
sqft2
or
LD50/
day3
0.5
Acute
Restricted
Use
EEC/
LC50
or
LD50/
sqft
or
LD50/
day
(
or
LD50
<
50
mg/
kg)
0.2
Acute
Endangered
Species
EEC/
LC50
or
LD50/
sqft
or
LD50/
day
0.1
Chronic
Risk
EEC/
NOEC
1
Table
4:
Risk
Presumptions
for
Terrestrial
Animals
Risk
Presumption
RQ
LOC
Page
13
of
65
Wild
Mammals
Acute
High
Risk
EEC/
LC50
or
LD50/
sqft
or
LD50/
day
0.5
Acute
Restricted
Use
EEC/
LC50
or
LD50/
sqft
or
LD50/
day
(
or
LD50
<
50
mg/
kg)
0.2
Acute
Endangered
Species
EEC/
LC50
or
LD50/
sqft
or
LD50/
day
0.1
Chronic
Risk
EEC/
NOEC
1
1
abbreviation
for
Estimated
Environmental
Concentration
(
ppm)
on
avian/
mammalian
food
items
2
mg/
ft2
3
mg
of
toxicant
consumed/
day
(
LD50
*
wt.
of
bird)

Table
5:
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
Table
6:
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
1.
Exposure
and
Risk
to
Nontarget
Terrestrial
Animals
Page
14
of
65
Terrestrial
organisms
may
be
exposed
to
CDDs/
CDFs
through
direct
contact
with
treated
lumber
and
contact
with
CDDs/
CDFs
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
CDDs/
CDFs
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
CDDs/
CDFs
leaching
occurs
from
treated
wood
into
the
surrounding
soil.
The
leached
compounds
in
soil
were
then
assumed
to
be
taken
up
by
terrestrial
vegetation.
Based
on
these
conditions,
the
model
assumed
that
birds
and
mammals
would
be
exposed
to
CDDs/
CDFs
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
the
"
Environmental
Fate
Modeling
RED
Chapter
for
Dioxin
in
Technical
Grade
Pentachlorophenol".
Using
the
state
of
Ohio
to
represent
a
typical
regional
environmental
setting,
U.
S.
EPA
calculated
the
amount
of
CDDs/
CDFs
released
from
an
estimated
1,500,000
utility
poles
within
the
state.
Based
on
the
calculated
release
rates
of
CDDs/
CDFs
via
volatilization,
wood
erosion,
and
leaching,
the
fate
and
transport
of
CDDs/
CDFs
releases
were
modeled
using
the
Level
III
model.
The
amounts
released
were
assumed
to
partition
into
the
air,
soil,
water,
and
sediments.
Amounts
of
individual
CDD
and
CDF
congeners
in
each
compartment
are
reported
in
the
Environmental
Fate
Modeling
chapter
of
this
document.
Since
the
toxicity
of
dioxins
and
dioxin­
like
compounds
is
known
to
be
additive
(
Tillit,
1999),
the
combined
concentration
of
CDDs/
CDFs
as
Toxic
Equivalents
(
TEQ)
was
used
as
the
EEC
in
this
risk
assessment,
as
well
as
the
concentration
of
TCDD
(
the
most
toxic
congener)
alone.
According
to
the
modeling
results,
terrestrial
organisms
would
be
exposed
to
2.32e­
02
mg/
kg
dioxin
TEQ
in
soil,
2.38E­
04
mg/
kg
of
which
is
due
to
TCDD..
Risk
Quotients
(
RQ
)
were
calculated
by
comparing
the
toxicity
endpoints
for
TCDD
to
the
TCDD
concentration,
as
well
as
the
TEQ
concentration.

The
methods
used
to
estimate
the
potential
dose
of
CDD/
CDF
to
avian
and
mammalian
species
through
the
ingestion
of
vegetation
and
soil
are
provided
in
Appendix
17.
A
summary
of
the
results
is
provided
in
the
table
below:

Table
7:
Summary
of
Risk
Quotients
for
Birds
and
Mammals
Species
Compound
Soil
Concentration
(
mg/
kg)
Total
Dose
(
EEC)
(
mg/
kgday
Endpoint
RQ
Bobwhite
CDD/
CDF
TEQ
2.32E­
02
9.49E­
04
TCDD
LD50=
1.5E­
05
mg/
kg
(
Hudson
et
al.
1984)
63.32a
TCDD
NOAEL
=
1.49E­
05
mg/
kg­
day
(
Schwetz
et
al.,
1973)
67.78b
TCDD
2.38E­
04
9.70E­
06
TCDD
LD50
=
1.5E­
05
mg/
kg
(
Hudson
et
al.
1984)
0.65c
Page
15
of
65
TCDD
NOAEL
=
1.49E­
05
mg/
kg­
day
(
Schwetz
et
al.,
1973)
0.69
Meadow
vole
CDD/
CDF
TEQ
2.32E­
02
3.57E­
03
TCDD
LD50
0.09
mg/
kg
(
Kociba
and
Schwetz,
1982,
adjusted
for
mouse/
vole
body
weight)
0.04
TCDD
NOAEL
=
9.94E­
04
(
Khera
and
Ruddick,
1973,
adjusted
for
rat/
vole
body
weight)
3.59b
TCDD
2.38E­
04
3.66E­
05
TCDD
LD50
0.09
mg/
kg
(
Kociba
and
Schwetz,
1982,
adjusted
for
mouse/
vole
body
weight)
4.07E­
04
TCDD
NOAEL
=
9.94E­
04
(
Khera
and
Ruddick,
1973,
adjusted
for
rat/
vole
body
weight)
0.04
a
Exceeds
the
Acute
High
Risk
LOC
b
Exceeds
the
Chronic
Risk
LOC
c
Exceeds
the
Restricted
Use
LOC
i.
Birds
As
shown
in
Table
7,
when
the
TEQ
concentrations
of
CDDs/
CDFs
are
used
to
determine
the
EEC,
the
resulting
RQs
exceed
the
LOC
for
acute
high
risk
and
chronic
risk.
When
the
single
TCDD
congener
soil
concentration
is
used
to
determine
the
EEC,
the
risk
is
lessened,
but
still
exceeds
the
Restricted
Used
LOC
for
acute
risk;
however,
using
only
the
TCDD
soil
concentration
to
address
exposure
may
underestimates
the
true
exposure
and
risk
to
terrestrial
mammals
from
CDDs/
CDFs
from
pentachlorophenol­
treated
structures,
as
discussed
below.

ii.
Mammals
Based
on
the
results
of
the
risk
assessment,
the
TEQ
soil
concentration
would
result
in
an
RQ
that
exceeds
the
chronic
LOC
for
mammals.
When
only
the
TCDD
soil
concentration
is
considered,
the
risk
is
lessened,
but
this
underestimates
the
true
exposure
and
risk
to
terrestrial
mammals
from
CDDs/
CDFs
from
pentachlorophenol­
treated
structures,
as
discussed
below.

iii.
Risk
Characterization
for
Birds
and
Mammals
Page
16
of
65
Using
only
the
TCDD
soil
concentration
to
estimate
exposure
to
CDDs/
CDFs
does
not
address
the
true
exposure
of
these
compounds.
As
described
the
Environmental
Fate
Modeling
chapter,
numerous
CDD/
CDF
congeners
will
be
present
in
soil
surrounding
pentachlorophenol­
treated
structures,
and,
while
TCDD
is
the
most
toxic
congener
to
birds
and
mammals,
it
is
present
in
lower
concentrations
than
many
other
congeners.
Bird
and
mammal
toxicity
data
for
CDDs/
CDFs
focuses
primarily
on
TCDD,
since
it
is
recognized
as
the
most
toxic
congener,
but
it
is
known
that
the
toxicity
of
CDDs/
CDFs
is
additive
(
Tillit,
1999).
Focusing
solely
on
TCDD
to
address
exposure
and
risk
would
therefore
likely
underestimate
the
risks
posed
to
terrestrial
wildlife
from
CDDs/
CDFs.
Even
comparing
the
TEQ
EEC
to
TCDD
endpoints
likely
underestimates
the
risk,
but
methods
to
extrapolate
TCDD
toxicity
to
be
representative
of
a
combined
dose
of
CDDs/
CDFs
were
not
available
at
the
time
of
this
assessment.
This
underestimation
may
be
offset,
at
least
in
part,
by
the
conservative
parameters
of
the
model
used.
The
model
assumes
that
the
entire
diet
of
the
bird
or
mammal
consists
of
soil
and
vegetation
from
the
immediate
vicinity
of
a
pentachlorophenol­
treated
structure,
and
thus
represents
a
"
worst­
case"
scenario.
In
reality,
the
foraging
area
of
the
animal
would
likely
include
areas
that
are
not
impacted
by
CDDs/
CDFs
released
from
a
treated
structure,
which
would
result
in
a
lower
actual
exposure
to
CDDs/
CDFs.
Data
and
models
to
quantify
this
were
not
available
at
the
time
of
this
assessment.

The
risk
assessment
for
birds
and
mammals
was
based
solely
on
exposure
to
CDDs/
CDFs
in
soil.
A
quantitative
assessment
of
the
risks
to
birds
and
mammals
from
direct
contact
with
the
treated
wood
was
not
conducted
due
to
the
lack
of
exposure
and
toxicity
data.
As
a
result,
the
potential
risks
from
direct
contact
with
pentachlorophenol­
treated
wood
were
not
evaluated.

Considering
the
uncertainties
discussed
above,
there
are
potential
acute
and
chronic
risks
to
birds
and
chronic
risks
to
mammals
from
CDDs/
CDFs
from
pentachlorophenol­
treated
wood,
especially
considering
the
tendency
of
CDDs/
CDFs
to
persist
and
bioaccumulate.
Sensitive
animals,
such
as
endangered
and
threatened
species,
may
be
particularly
at
risk.
There
are
also
CDDs/
CDFs
present
in
the
environment
from
other
sources,
and
those
contributed
by
the
use
of
pentachlorophenol
are
adding
to
the
total
environmental
load
and
subsequent
risk
of
these
compounds
B.
Exposure
and
Risk
to
Fish
and
Aquatic
Invertebrates
Fish
and
aquatic
invertebrates
may
potentially
be
exposed
to
CDDs/
CDFs
leached
from
utility
poles
into
surrounding
soil,
which
then
enters
aquatic
habitats
via
runoff.
Most
(>
99
%)
CDD/
CDF
will
partition
into
the
sediment,
but
a
small
amount
will
remain
in
the
water
column.
Peak
concentrations
of
TCDD/
TCDF
(
from
PRZM/
EXAMS
modeling)
are
presented
in
Table
8,
below.
Detailed
descriptions
of
the
models
used
to
predict
these
EECs
can
be
found
in
the
Environmental
Fate
Modeling
chapter
of
this
document.

Table
8:
Peak
concentrations
(
36
year)
of
TCDD,
TCDF
and
TEQ
in
various
aquatic
media,
as
predicted
by
PRZM­
3
and
EXAMS
models
(
excerpted
from
Environmental
Modeling
chapter
of
this
document)
Page
17
of
65
2,3,7,8­
TCDD
TEQ
Water
column,
total
(
ng/
L)
3.68E­
8
2.69E­
6
Water
column,
dissolved
(
ng/
L)
2.77E­
9
6.45E­
8
Water
column
sediment
(
ng/
L)
1.10E­
6
8.53E­
5
Sediment
total
(
ng/
L)
1.13E­
6
1.23E­
4
Sediment,
pore
water
diss.
(
ng/
L)
2.83E­
9
1.74E­
7
Sediment,
benthic
(
ng/
g)
1.13E­
6
1.23E­
4
Sediment,
benthic
organism
(
ng/
g)
1.81E­
6
1.66E­
4
Fish
(
ng/
g­
lipid)
1.02E­
6
1.73E­
5
The
peak
total
water
column
concentration
of
2,3,7,8­
TCDD
was
compared
to
the
aquatic
toxicity
endpoints
provided
in
Table
1.
Since
the
toxicity
of
the
various
congeners
of
dioxins
and
furans
is
known
to
be
additive
(
Tillitt,
1999),
the
total
concentration
(
as
TEQ)
was
also
compared
to
the
2,3,7,8­
TCDD
endpoint;
however,
this
may
still
underestimate
the
risk
posed
by
CDDs/
CDFs
since
the
toxicity
endpoint
is
for
only
the
most
sensitive
congener,
and
not
a
combination
of
congeners.

Table
9:
TCDD
and
TEQ
Risk
Quotients
for
Aquatic
Organisms
Organism
class
Species
Endpoint
type
and
value
RQ
TCDD
RQ
TEQ
Freshwater
fish
Rainbow
trout,
Oncorhynchus
mykiss
96­
hour
acute
=
1.83
ng/
L
(
Bol
et
al.,
1989)
2.01
E­
08
1.40E­
06
Freshwater
fish
Rainbow
trout,
Oncorhynchus
mykiss
28­
day
NOEC
=
1.1E­
3
ng/
L
(
Mehrle
et
al.,
1988)
3.34
E­
05
2.44E­
03
Freshwater
invertebrate
Water
flea,
Daphnia
magna
48­
hour
acute
EC50
=
1020
ng/
L
(
Adams
et
al.,
1986)
3.61
E­
11
2.64E­
09
Freshwater
invertebrate
Snail
(
exact
species
not
reported)
12­
day
reproductive
NOEC
=
200
ng/
L
(
Miller
et
al.,
1973)
1.84
E­
10
1.34E­
08
All
of
these
are
well
below
any
Level
of
Concern
for
aquatic
organisms.

CDDs/
CDFs
which
are
bound
to
sediment
can
be
taken
up
by
benthic
organisms,
which
in
turn
are
consumed
by
fish
and
higher­
order
aquatic
invertebrates.
The
bioaccumulative
tendencies
of
CDDs/
CDFs
could
result
in
their
concentrations
magnifying
through
the
food
chain,
eventually
reaching
toxic
levels.
The
likelihood
of
this
occurring
with
CDD/
CDF
residues
resulting
from
the
use
of
pentachlorophenol
is
low,
since
the
initial
levels
of
CDDs/
CDFs
entering
soil,
water
and
sediment
are
well
below
reported
toxic
levels.
However,
there
are
CDDs/
CDFs
present
in
the
Page
18
of
65
environment
from
other
sources,
and
those
contributed
by
the
use
of
pentachlorophenol
are
adding
to
the
total
environmental
load
of
these
compounds.

C.
Exposure
and
Risk
to
Plants
There
are
no
terrestrial
or
aquatic
plant
endpoints
available
to
use
in
a
risk
assessment.
A
single
dibenzofuran
EC50
value
of
1500
ppb
for
the
diatom
Skeletonema
costatum
was
found
in
the
ACQUIRE
database.
Since
the
endpoints
for
other
aquatic
organisms
indicate
that
they
are
more
sensitive
than
aquatic
plants
to
dibenzofuran,
it
is
assumed
that
the
risk
assessment
for
fish
and
aquatic
invertebrates
is
sufficient
to
protect
aquatic
plant
species
as
well.

D.
Ecological
Risk
Conclusions
Based
on
the
quantitative
estimates
provided
above,
the
acute
and
short­
term
chronic
risk
to
fish,
aquatic
organisms
and
terrestrial
wildlife
from
pentachlorophenol­
related
CDDs/
CDFs
appears
to
be
low.
There
is
a
great
deal
of
uncertainty
in
these
quantitative
estimates,
however,
due
to
a
lack
of
standard
toxicity
endpoints
and
exposure
information.
There
are
also
uncertainties
and
limitations
inherent
to
the
exposure
models
used,
as
described
in
the
Environmental
Modeling
chapter
of
this
document.
Based
on
the
information
that
is
available,
it
is
unlikely
that
the
CDDs/
CDFs
resulting
from
the
use
of
pentachorophenol
on
utility
poles
are
posing
an
immediate
risk
to
wildlife
or
aquatic
organisms;
however,
it
is
possible
that
the
buildup
of
these
persistent
and
bioaccumulative
compounds
over
time
may
eventually
reach
levels
that
pose
a
risk
to
these
organisms.
Additionally,
pentachlorophenol
is
only
one
of
many
sources
of
CDDs/
CDFs
in
the
environment;
it
is
not
currently
possible
to
quantify
the
portion
of
the
aggregate
environmental
risk
from
CDDs/
CDFs
that
is
attributable
to
the
use
of
pentachlorophenol
on
utility
poles.
Page
19
of
65
VII.
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R.;
Seinen,
W.;
van
den
berg,
M.
(
1992)
Concurrence
of
P450
1A1
induction
and
toxic
effects
after
administration
of
a
low
dose
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
(
TCDD)
to
the
rainbow
trout
(
Oncorhynchus
mykiss).
Aquat.
Toxicol.
24:
123­
142.

Walker,
M.
K.
(
1991)
Toxicity
of
polychlorinated
dibenzo­
p­
dioxins,
polychlorinated
dibenzofurans,
and
polychlorinated
diphenyls
during
salmonid
early
development.
Ph.
D.
thesis.
University
of
Wisconsin,
Madison,
WI,
August
1991.

Walker,
M.
K.;
Spitsbergen,
J.
M.;
Olson,
J.
R.;
Peterson,
R.
E.
(
1991)
2,3,7,8­
tetrachlorodibenzop
dioxin
(
TCDD)
Toxicity
during
early
life
stage
development
of
lake
trout
(
Salvelinus
namaycush).
Can.
J.
Fish.
Aquat.
Sci.
48(
5):
875­
883.
In:
Prog.
Abstr.
32nd
Conf.
Int.
Assoc.
Great
Lakes
Res.,
May
30­
June
2,
1989,
Univ.
of
Wisconsin,
Madison,
WI:
114(
ABS).

Walker,
M.
K.;
Peterson,
R.
E.
(
1991)
Potencies
of
polychlorinated
dibenzo­
p­
dioxin,
dibenzofuran,
and
biphenyl
congeners,
relative
to
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin,
for
producing
early
life
stage
mortality
in
rainbow
trout
(
Oncorhynchus
mykiss).
Aquat.
Toxicol.
21:
219­
238.

Walker,
M.
K.;
Spitsbergen,
J.
M.;
Olson,
J.
R.;
Peterson,
R.
E.
(
1991)
2,3,7,8­
tetrachlorodibenzop
dioxin
(
TCDD)
toxicity
during
early
life
stage
development
of
lake
trout
(
Salvenlinus
namaycush).
Can.
J.
Fish.
Aquat.
Sci.
48:
875­
883.

Walker,
M.
K.;
Hufnagle,
Jr.,
L.
C.;
Clayton,
M.
K.;
Peterson,
R.
E.
(
1992a)
An
egg
injection
method
for
assessing
early
life
stage
mortality
of
polychlorinated
dibenzo­
p­
dioxins,
dibenzofurans,
and
biphenyls
in
rainbow
trout...
Aquat.
Toxicol.
22:
15­
38.

Walker,
M.
K.;
Hufnagle,
Jr.
L.
C.;
Clayton,
M.
K.;
Peterson,
R.
E.
(
1992)
An
egg
injection
method
for
assessing
early
life
stage
mortality
of
polychlorinated
dibenzo­
p­
dioxins,
dibenzofurans,
and
biphenyls
in
rainbow
trout,
(
Oncorhynchus
mykiss).
Aquat.
Toxicol.
22:
15­
38.

Walker,
M.
K.
(
1992)
Toxicity
of
polychlorinated
dibenzo­
p­
dioxins,
polychlorinated
dibenzofurans,
and
polychlorinated
biphenyls
during
salmonid
early
development.
Diss.
Abstr.
Int.
B
Sci.
Eng.
52(
10:
5177
(
ABS).

Walker,
M.
K.;
Cook,
P.
M.;
Batterman,
A.
R.;
Lothenbach,
D.
B.;
Berini,
C.;
Hufnagle,
L.;
Peterson,
R.
E.
(
1993)
Early
life
stage
mortality
associated
with
maternal
transfer
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
to
lake
trout
oocytes.
U.
S.
EPA.
Environmental
Research
Laboratory,
Duluth,
MN.
(
In
preparation).
Page
27
of
65
Wannemacher,
R.;
Rebstock,
A.;
Kulzer,
E.;
Schrenk,
D.;
Bock,
K.
W.
(
1992)
Effects
of
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
on
reproduction
and
oogenesis
in
zebrafish
(
Brachydanio
rerio).
Chemosphere
24:
1361­
1368.

Wisk,
J.;
Cooper,
K.
R.
(
1986)
Comparison
of
toxicity
between
2,3,7,8­
tetrachlorodibenzo­
pdioxin
(
TCDD)
and
several
tetrachlorodibenzofuran
isomers
(
TCDF)
in
the
Japanesa
medaka
embryo­
larval
bioassay.
Proceedings
of
the
7th
Annual
Meeting
of
SETAC.

Wisk,
J.
D.;
Cooper,
K.
R.
(
1990a)
Comparison
of
the
toxicity
of
several
polychlorinated
dibenzop
dioxins
and
2,3,7,8­
tetrachlorodibenzofuran
in
embryos
of
the
Japanese
medaka
(
Oryzias...).
Chemosphere
20(
3­
4):
361­
377.

Wisk,
J.
D.;
Cooper,
K.
R.
(
1990b)
The
stage
specific
toxicity
of
2,3,7,8­
tetrachlorodibenzo­
pdioxin
in
embryos
of
the
Japanese
medaka
(
Oryzias
latipes).
Environ.
Toxicol.
Chem.
9:
1159­
1169.

Yalkowsky,
S.
H.;
Valvani,
S.
C.;
Mackay,
D.
(
1983)
Estimation
of
the
aqueous
solubility
of
some
aromatic
compounds.
Residue
Rev.
85:
43­
55.

Yockum,
R.
S.;
Isensee,
A.
R.;
Jones,
G.
E.
(
1978)
Distribution
and
toxicity
of
TCDD
and
2,4,5­
T
in
an
aquatic
model
ecosystem.
Chemosphere
7:
215­
220.

Zabel,
E.
W.;
Walker,
M.
K.;
Hornung,
M.
W.;
Clayton,
M.
K.;
Peterson,
R.
E.
(
1995)
Interactions
of
polychlorinated
dibenzo­
p­
dioxin,
dibenzofuran,
and
biphenyl
congeners
for
producing
rainbow
trout
early
life
stage
mortality.
Toxicol.
Appl.
Pharmacol.
134:
204­
213.

Zabel,
E.
W.;
Cook,
P.
M.;
Peterson,
R.
E.
(
1995b)
Toxic
equivalency
factors
of
polychlorinated
dibenzo­
p­
dioxin,
dibenzofuran
and
biphenyl
congeners
based
on
early
life
stage
mortality
in
rainbow
trout.
Aquat.
Toxicol.
31(
4):
315­
328.

Zabel,
E.
W.;
Cook,
P.
M.;
Peterson,
R.
E.
(
1995c)
Potency
of
3,3',
4,4',
5­
pentachlorobiphenyl
(
PCB
126),
alone
and
in
combination
with
2,3,7,8­
tetrachlorodibenzo­
p­
dioxin
(
TCDD),
to
produce
lake
trout.
Environ.
Toxicol.
Chem.
14(
12):
2175­
2179.

Zabel,
E.
W.;
Peterson,
R.
E.
(
1996)
TCDD­
like
activity
of
2,3,6,7­
tetrachloroxanthene
in
rainbow
trout
early
life
stages
and
in
a
rainbow
trout
gonadal
cell
line
(
RTG­
2).
Environ.
Toxicol.
Chem.
15(
12):
2305­
2309.
Page
28
of
65
Appendices:
Toxicity
Data
for
Terrestrial
and
Aquatic
Organisms
Page
29
of
65
Appendix
1.
Compilation
of
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzo­
p­
Dioxins
(
CDDs)
(
U.
S.
EPA,
1990)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Monochloro­
No
Data
Dichloro­
No
Data
Trichloro­
No
Data
Tetrachloro­

1,2,3,7­
TCDD
Rainbow
Trout
(
fry
­
0.2
g)
5
(
water)
+
24
Days
134
54
No
deaths
No
deaths
Muir
et
al.
(
1985)

Fathead
Minnow
(
0.2
g)
5
(
water)
+
24
Days
28
23
No
deaths
No
deaths
Muir
et
al.
(
1985)

Rainbow
Trout
(
0.5
­
1.0g)
30
(
diet)
+
30
Days
110
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)

Fathead
Minnow
(
1.0
g)
30
(
diet)
+
30
Days
110
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)

1,3,6,8­
TCDD
Rainbow
Trout
(
0.1
­
0.5
g)
5
(
water)
+
48
Days
+
48
Days
+
48
Days
211
74
4
No
deaths
No
deaths
No
deaths
Muir
et
al.
(
1986)

Fathead
Minnow
(
1.5
­
2.5
g)
5
(
water)
+
29
Days
+
48
Days
41
10
No
deaths
No
deaths
Muir
et
al.
(
1986)

2,3,7,8­
TCDD
Fathead
Minnow
20
(
water)
+
20
Days
1.0
3
out
of
30
dead
Adams
et
al.
(
1986)

Fathead
Minnow
1,2,3,4
(
water
+
60
Days
0.0
0.12
10%
dead
<
5%
dead
Adams
et
al.
(
1986)
Appendix
1.
Compilation
of
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzo­
p­
Dioxins
(
CDDs)
(
U.
S.
EPA,
1990)
(
continued)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Page
30
of
65
Fathead
Minnow
28
(
water)
+
20
Days
1.7
6.7
63
53%
dead
(
28
days)

100%
dead
(
22
days)

100%
dead
(
12
days)

28­
d
LC50
1.7
ng/
L
Adams
et
al.
(
1986)

2,3,7,8­
TCDD
(
continued)
Coho
Salmon
1
(
water)
0.056
Deaths
after
several
weeks
Miller
et
al.
(
1973)

Rainbow
Trout
(
35
g)
0.25
(
water)
+
139
Days
107
4
Deaths,
one
each
on
Days
78,
136,
137,
139
and
decreased
growth
Branson
et
al.
(
1985)

Rainbow
Trout
(
yolk
sac
fry)
4
(
water)
+
16
Days
54.6
33.1
20.1
7.4
4.5
2.7
1.6
100%
dead
(
13
days)

100%
dead
(
15
days)

100%
dead
(
17
days)

95%
dead
(
20
days)

63%
dead
(
20
days)

20%
dead
(
20
days)

4%
dead
(
20
days)
Helder
and
Seinen
(
1985
Rainbow
Trout
(
fry
­
0.38
g)
28
(
water)
+
28
Days
0.789
0.382
0.176
0.079
0.038
0.0011
(
control)
85%
dead
(
28
days)

73%
dead
(
28
days)

50%
dead
(
28
days)

95%
dead
(
56
days)

18%
dead
(
28
days)

83%
dead
(
56
days)

6%
dead
(
28
days)

45%
dead
(
56
days)

5%
dead
(
28
days)

7%
dead
(
56
days)
Mehrle
et
al.
(
1988)

Daphnia
magna
2
(
water)
+
7
Days
0.2
­
1,030
No
mortality
48­
hr
EC50
>
1,030
ng/
L
Adams
et
al.
(
1986)
Appendix
1.
Compilation
of
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzo­
p­
Dioxins
(
CDDs)
(
U.
S.
EPA,
1990)
(
continued)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Page
31
of
65
Rainbow
Trout
91
(
diet)
+
91
Days
494
No
overt
effects
Kleeman
et
al.
(
1986a)

Yellow
Perch
91
(
diet)
+
91
Days
494
No
overt
effects
Kleeman
et
al.
(
1986b)

2,3,7,8­
TCDD
(
continued)
Rainbow
Trout
1
dose
intraperitoneal
injection
(
i.
p.)
+
14
Days
10
ng/
kg
1.0
0.1
0.01
Hypophagic,
etc.

Normal
Normal
Normal
Spitsbergen
(
1986)

Rainbow
Trout
(
LC50
­
10

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
95%
dead
(
80
days)

90%
dead
(
80
days)

20%
dead
(
80
days)

0%
dead
(
80
days)
Spitsbergen
et
al.
(
1988a)

Rainbow
Trout
(
LD50
­
3

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
95%
dead
(
22
days)

100%
dead
(
40
days)

80%
dead
(
75
days)

8%
dead
(
80
days)
Kleeman
et
al.
(
1986a)

Carp
(
LD50
­
3

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
100%
dead
(
45
days)

100%
dead
(
55
days)

90%
dead
(
75
days)

5%
dead
(
80
days)
Kleeman
et
al.
(
1988)

Bullhead
(
LD50
­
5

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
100%
dead
(
22
days)

100%
dead
(
22
days)

50%
dead
(
65
days)

3%
dead
(
80
days)
Kleeman
et
al.
(
1988)
Appendix
1.
Compilation
of
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzo­
p­
Dioxins
(
CDDs)
(
U.
S.
EPA,
1990)
(
continued)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Page
32
of
65
Largemouth
Bass
(
LD50
­
11

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
100%
dead
(
22
days)

95%
dead
(
80
days)

5%
dead
(
80
days)

5%
dead
(
80
days)
Kleeman
et
al.
(
1988)

Bluegill
(
LD50
­
16

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
100%
dead
(
23
days)

80%
dead
(
76
days)

2%
dead
(
80
days)

5%
dead
(
80
days)
Kleeman
et
al.
(
1988)

2,3,7,8­
TCDD
(
continued)
Yellow
Perch
(
LD50
­
10

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
100%
dead
(
35
days)

90%
dead
(
40
days)

20%
dead
(
60
days)

5%
dead
(
80
days)
Kleeman
et
al.
(
1986b)

Yellow
Perch
(
LC50
­
3

g/
kg)
1
dose
(
i.
p.)
+
80
Days
125
25
5
1
<
95%
dead
(
28
days)

<
95%
dead
(
28
days)

80%
dead
(
80
days)

0%
dead
(
80
days)
Spitsbergen
et
al.
(
1988b)

Pentachloro­

1,2,3,4,7­
PCDD
Rainbow
Trout
(
fry
­
0.2
g)
5
(
water)
+
24
Days
16
No
deaths
Muir
et
al.
(
1985)

Fathead
Minnow
(
0.2
g)
5
(
water)
+
24
Days
+
24
Days
19
11
No
deaths
No
deaths
Muir
et
al.
(
1985)

Rainbow
Trout
(
0.5
­
1.0
g)
30
(
diet)
+
30
Days
105
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)

Fathead
Minnow
(
1.0
g)
30
(
diet)
+
30
Days
105
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)
Appendix
1.
Compilation
of
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzo­
p­
Dioxins
(
CDDs)
(
U.
S.
EPA,
1990)
(
continued)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Page
33
of
65
Hexachloro­

1,2,3,4,7,8­
HxCDD
Rainbow
Trout
(
fry
­
0.2g)
5
(
water)
+
24
Days
47
26%
dead
(
12
days)
and
weight
loss
Muir
et
al.
(
1985)

Rainbow
Trout
(
fry
­
0.2g)
5
(
water)
+
48
Days
10
No
effects
Muir
et
al.
(
1985)

Fathead
Minnow
(
0.2
g)
5
(
water)
+
24
Days
+
48
Days
18
7
No
deaths
No
deaths
Muir
et
la.
(
1985)

1,2,3,4,7,8­
HxCDD
(
continued)
Rainbow
Trout
(
0.5
­
1.0
g)
30
(
diet)
+
30
Days
109
ng/
g
Significant
(
P=
0.05)

growth
rate
effects
Muir
et
al.
(
1988)

Fathead
Minnow
(
1.0
g)
30
(
diet)
+
30
Days
109
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)

1,2,3,7,8,9­
HxCDD
Rainbow
Trout
(
yolk
sac
fry)
4
(
water)
+
10
Days
Less
than
10­
5
times
as
toxic
as
2,3,7,8­
TCDD
Helder
and
Seinen
(
1986)

Heptachloro­
1,2,3,4,6,7,8­
HpCDD
Rainbow
Trout
(
fry
­
0.2
g)
5
(
water)
+
24
Days
+
48
Days
55
11
No
deaths
No
deaths
Muir
et
al.
(
1985)

Fathead
Minnow
(
0.2
g)
5
(
water)
+
24
Days
+
48
Days
39
8
No
deaths
No
deaths
Muir
et
al.
(
1985)

Rainbow
Trout
(
0.5
­
1.0
g)
30
(
diet)
+
30
Days
109
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)

Fathead
Minnow
(
1.0
g)
30
(
diet)
+
30
Days)
109
ng/
g
No
significant
(
P=
0.05)

adverse
effects
Muir
et
al.
(
1988)

Octachloro­
Appendix
1.
Compilation
of
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzo­
p­
Dioxins
(
CDDs)
(
U.
S.
EPA,
1990)
(
continued)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Page
34
of
65
Rainbow
Trout
(
0.1
­
0.5
g)
5
(
water)
+
18
Days
+
32
Days
415
20
No
deaths
No
deaths
Muir
et
al.
(
1986)

Fathead
Minnow
(
1.5
­
2.5
g)
5
(
water)
+
48
Days
9
No
deaths
Muir
et
al.
(
1986)

Source:
U.
S.
EPA,
1990.
Page
35
of
65
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
U.
S.
EPA,
1993)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
AQUATIC
LIFE
Freshwater
Species
Algae,
Oedogonium
cardiacum
Model
ecosystem
1,330
2,295,000
33­
d
No
toxic
effect
Isensee
and
Jones,
1975;

Isensee,
1978
Vascular
plant
Duckweed,

Lemna
minor
Model
ecosystem
1,330
33­
d
No
toxic
effect
Isensee
and
Jones,
1975;

Isensee,
1978
7.13
30,700
33­
d
No
toxic
effect
Isensee
and
Jones,
1975
Annelid
Worm,
Paranais
sp.
Water
(
static)
200c
55­
d
No
decrease
in
reproductive
success
Miller
et
al.,
1973
Mollusc
Snail
(
adult),

Physa
sp.
Model
ecosystem
1,330
502,000
33­
d
No
toxic
effect
Isensee
and
Jones,
1975;

Isensee,
1978
Water
(
static)
200c
36­
d
12­
d
No
decrease
in
reproductive
success
Miler
et
al.,
1973
Arthropod
Mosquito
(
larvae),

Aedes
aegypti
Water
(
static)
200c
17­
d
23­
d
No
effect
on
pupation
Miller
et
al.,
1973
Cladoceran
(
adult),

Daphnia
magna
Model
ecosystem
1,330
1,570,000
33­
d
No
toxic
effect
Isensee
and
Jones,
1975
Cladoceran
(
1­
21
d),

Daphnia
magna
Water
(
renewal)
1,030
48­
h
7­
d
No
toxic
effect
Adams
et
al.,
1986
Fish
Coho
salmon,
Oncorhynchus
kisutch
Juvenille
(
3.5
g)
Water
(
static)
0.56
96­
h
114­
d
No
toxic
effect
Miller
et
al.,
1979
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Page
36
of
65
5.60
96­
h
56­
d
50%
mortality
Miller
et
al.,
1973,
1979
Rainbow
trout,

Oncorhynchus
mykiss
Eggs
Water
(
renewal)
0.10c
96­
h
160­
d
Delayed
development,

reduced
growth
of
fry
Helder
1981;
1982a,
b
1c
96­
h
160­
d
Reduced
growth,

mortality
in
sac
fry
Helder,
1981;
1982a,
b
10c
96­
h
40­
d
100%
mortality
in
sac
fry
Helder,
1982a,
b
Egg
injection
230
(
in
eggs)
Single
injection
Fertilized
egg
to
swim­
up
LR50
(
sac
fry
of
McConaughy
strain)
Walker
and
Peterson,

1991
240
(
in
eggs)
Single
injection
Fertilized
egg
to
swim­
up
LR50
(
sac
fry
of
Erwin
strain)
Walker
and
Peterson,

1991
374
(
in
eggs)
Single
injection
Fertilized
egg
to
swim­
up
LR50
(
sac
fry
of
Arlee
strain)
Walker
and
Peterson,

1991
488
(
in
eggs)
Single
injection
Fertilized
egg
to
swim­
up
LR50
(
sac
fry
of
Eagle
Lake
strain)
Walker
and
Peterson,

1991
421
(
in
eggs)
Single
injection
>
48­
h
to
post
swim­
up
LR50
(
sac
fry
of
Fish
Lake
strain)
Walker
et
al.,
1992
Water
(
renewal)
279
(
in
eggs)
48­
h
>
48­
h
to
post
swim­
up
Significant
mortality
in
sac
fry
Walker
et
al.,
1992
439
(
in
eggs)
48­
h
>
48­
h
to
post
swim­
up
LR50d
(
sac
fry)
Walker
et
al.,
1992
Sac
fry
Water
(
renewal)
1c
96­
h
160­
d
Reduced
growth,

mortality
Helder
1981;
1982a,
b
10c
96­
h
10­
d
100%
mortality
Helder
1982a,
b
12.2c
96­
h
16­
d
100%
mortality
Helder
and
Seinen,
1985
1.83c
96­
h
21­
d
LC50
Bol
et
al.,
1989
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Page
37
of
65
Swim­
up
fry
(
0.38
g)
Water
(
flow­
thru)
0.176
3,220
28­
d
28­
d
95%
mortality
Mehrle
et
al.,
1988
0.0011
21g
28­
d
28­
d
NOAELe
Mehrle
et
al.,
1988
0.038
765g
28­
d
28­
d
LOAELf
(
45%
mortality)
Mehrl
et
al.,
1988
0.046
28­
d
28­
d
LC50
Mehrle
et
al.,
1988
Juvenile
(
0.85
g)
Water
(
renewal)
10
96­
h
68­
d
Reduced
growth,

mortality
Helder
1981;
1982a,
b
100
96­
h
23­
d
100%
mortality
Helder
1981;
1982a,
b
Fingerling
(
25­
45
g)
i.
p.
injection
1,000c
Single
injection
25­
d
Significant
hematological
changes
Spitsbergen
et
al.,
1988a
5,000c
Single
injection
20­
d
20%
mortality
Spitsbergen
et
al.,
1988a
5,000c
Single
injection
11­
12­
wk
20%
mortality,
increased
liver
weight
van
der
Weiden
et
al.,

1990
10,000c
Single
injection
80­
d
LD50
Spitsbergen
et
al.,

1988a;
Kleeman
et
al.,

1988
Fingerling
(
35
g)
Water
(
static)
107
650­
2,580
6­
h
42­
139­
d
Mortality,
fin
rot,

increased
liver
weight
Branson
et
al.,
1985
Fingerling
(
7.8
cm)
Diet
(
3.29
ng/
g)
314
71­
d
No
effect
on
survival
and
growth
Hawkes
and
Norris,

1977
Diet
(
1,700
ng/
g)
276,000
71­
d
100%
mortality
Hawkes
and
Norris,

1977
Fingerling
(
3­
7
g)
Diet
(
0.494
ng/
g)
250
13­
wk
13­
wk
No
toxic
effect
Kleeman
et
al.,
1986a
Fingerling
(
8
g)

Yearling
(
100­
150
g)
i.
p.
injection
10,000c
Single
injection
2­
4
wk
post
exposure
Fin
necrosis,
no
effect
on
immune
suppression
Spitsbergen
et
al.,
1986;

1988c
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Page
38
of
65
Juvenile
(
46
g)
i.
p.
injection
300­
3,060
Single
injection
6­
12
wk
Fin
hemorrhage,
spleen
histopathology,
EROD
induction,
P4501A1
induction
van
der
Welden
et
al.,

1992
790
Single
injection
3­
wk
ED50
for
EROD
induction
van
der
Welden
et
al.,

1992
Immature
Adult
(
300­

400
g)
i.
p.
injection
640c
Single
injection
72­
h
ED50
for
AHH
induction
Janz
and
Metcalfe,
1991
Lake
trout,
Salvelinus
namaycush
Eggs
Water
(
renewal)
34
(
in
eggs)
48­
h
>
48­
h
to
post
swim­
up
NOAEL
Walker
et
al.,
1991
40
(
in
eggs)
48­
h
>
48­
h
to
post
swim­
up
23%
mortality
in
sac
fry
Spitsbergen
et
al.,
1991
55
(
in
eggs)
48­
h
>
48­
h
to
post
swim­
up
LOAEL
(
sac
fry
mortality)
Walker
et
al.,
1991
65
(
in
eggs)
48­
h
>
38­
h
to
post
swim­
up
LR50
(
sac
fry)
Walker
et
al.,
1991
Egg
injection
47
(
in
eggs)
Single
injection
Fertilized
egg
to
swim­
up
fry
LR50
(
sac
fry)
Walker
et
al.,
1992
Adult
Dieth
59
(
in
eggs)
90­
d
Eggs
to
swimup
fry
LR50
(
sac
fry)
Walker,
1991;
Walker
et
al.,
1993
104
(
in
eggs)
90­
d
Eggs
to
swimup
fry
100%
mortality
to
sac
fry
Walker,
1991;
Walker
et
al.,
1993
Northern
pike,
Esox
lucius
Eggs
Water
(
renewal)
0.1c
96­
h
72­
d
Delayed
hatch,
reduced
growth
of
fry
Helder,
1980;
1982a,
b
1.0
96­
h
53%
mortality
to
fry
Helder,
1980;
1982a,
b
10.0
96­
h
99%
mortality
to
fry
Helder,
1980;
1982a,
b
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Page
39
of
65
Carp,
Cyprinus
carpio
Juvenile
(
20
g)
i.
p.
injection
3,000c
Single
injection
80­
d
LD50
Kleeman
et
al.,
1988
Adult
Water
(
flow­
thru)
0.060
2,200
71­
d
61­
d
Mortality
and
pathology
Cook
et
al.,
1991
Zebrafish,
Brachydanio
revio
Adult
Diet
(
1.7
ng/
g)
Single
dose
22­
d
No
effect
Wannemacher
et
al.,

1992
Diet
(

8.3
ng/
g)
Single
dose
1­
2
Spawnings
Reduced
eggs
per
spawn,

100%
larval
mortality
Wannemacher
et
al.,

1992
Fathead
minnow,
Pimephales
promelas
Juvenile
(
1.0
g)
Water
(
flow­
thru)
0.049­

0.067
71­
d
61­
d
Mortality
and
pathology
Cook
et
al.,
1991
1.7
28­
d
LC50
Adams
et
al.,
1986
Juvenile
(
1.0­
2.0
g)
Water
(
static)
7.1
24­
h
60­
d
40%
mortality
Adams
et
al.,
1986
Bullhead,
Ictalurus
melas
Juvenile
(
6
g)
i.
p.
injection
5,000c
Single
injection
80­
d
LD50
Kleeman
et
al.,
1988
Channel
catfish,
Ictalurus
punctatus
Fingerling
Model
ecosystem
2.4­
4.2
15­
20­
d
100%
mortality
Yockim
et
al.,
1978
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Page
40
of
65
Japanese
medaka,
Oryzias
latipes
Eggs
S,
M
3.5­
6.0
Fertilized
egg
to
3­
d
post
hatch
EC50
(
embryos
with
lesions)
Wisk
and
Copper,

1990a,
b
9.0­
13.0
Fertilized
egg
to
3­
d
post
hatch
LC50
Wisk
and
Cooper,

1990a,
b
14.0­
15.0
Fertilized
egg
to
3­
d
post
hatch
EC50
(
embryos
with
severe
lesions)
Wisk
and
Cooper,

1990a,
b
14.0­
17.0
Fertilized
egg
to
3­
d
post
hatch
EC50
(
prevent
hatch)
Wisk
and
Cooper,

1990a,
b
Water
(
static)
240
(
in
embryos)
Fertilized
egg
to
3­
d
post
hatch
ER50i
(
embryos
with
lesions)
Wisk
and
Cooper,
1990b
Mosquito
fish,

Gambusia
affinis
Model
ecosystem
2.4­
4.2
15­
d
100%
mortality
Yockim
et
al.,
1978
Guppy,
Poecilia
reticulata
Juvenile
(
8­
12)
Water
(
static)
0.1c
24­
h
42­
d
Significant
increase
in
fin
necrosis
Miller
et
al.,
1979
Juvenile
(
9­
40
mm)
Water
(
static)
100c
120­
h
37­
d
100%
mortality
Norris
and
Miller,
1974
Bluegill,
Lepomis
macrochirus
Juvenile
(
30
g)
i.
p.
injection
16,000c
Single
injection
80­
d
LD50
Kleeman
et
al.,
1988
Largemouth
bass,
Micropterus
salmoides
Juvenile
(
7
g)
i.
p.
injection
11,000c
Single
injection
80­
d
LD50
Kleeman
et
al.,
1988
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Page
41
of
65
Yellow
perch,
Perca
flavescens
Juvenile
(
3­
6
g)
Diet
(
0.494
ng/
g)
143
13­
wk
13­
wk
No
toxic
effect
Kleeman
et
al.,
1986b
Juvenile
i.
p.
injection
3,000c
Single
injection
80­
d
LD50
Spitsbergen
et
al.,

1988b;
Kleeman
et
al.,

1988
Amphibian
Bullfrog,
Rana
catesbeiana
Tadpole
i.
p.
injection
1,000,000c
Single
injection
50­
d
No
effect
on
metamorphosis
Beatty
et
al.,
1976
Adult
i.
p.
injection
500,000c
Single
injection
35­
d
No
toxic
effect
Beatty
et
al.,
1976
Saltwater
Species
Rays
Little
Skate,
Raja
erinacea
500­
1,100
g
i.
p.
injection
1,000c
Single
injection
10­
d
Increased
enzyme
activity
Bend
et
al.,
1974
4,500c
Two
injections
7­
12­
d
10­
fold
increase
in
enzyme
activity
Pohl
et
al.,
1974
Fish,
Mummichog,
Fundulus
heteroclitus
Eggs
Water
(
static)
200c
Fertilized
egg
to
hatch
20%
mortality
and
50%

lesions
in
embryos
Cooper,
1989;
Prince
and
Cooper,
1989
Winter
flounder,
Pleuronectes
americanus
250
g
Oral
dose
4,500c
Two
doses
8­
d
Increased
enzyme
activity
Pohl
et
al.,
1974
a
Measured
TCDD
concentration
in
water.

b
Measured
TCDD
concentration
in
organism
(
wet
weight).

c
Unmeasured
TCDD
concentration
in
water
or
organism
(
wet
weight).

d
LR50
(
corrected
for
control
mortality)
term
defined
in
this
report
as
the
measured
residue
concentration
in
eggs
that
caused
50%
mortality
to
sac
fry.

e
NOAEL
=
No
observed
adverse
effect
level.

f
LOAEL=
Lowest
observed
adverse
effect
level.
Appendix
2.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Aquatic
Life
(
continued)

Page
42
of
65
g
NOAEL
and
LOAEL
values
(
based
on
mean
measured
wet
weight
organism
concentrations)
were
calculated
for
this
report.

h
Diet
consisted
of
22
ng/
g
pelletized
feed
followed
by
fathead
minnows
injected
with
500
pg/
fish.

i
ER50
­
Term
defined
in
this
report
as
the
measured
residue
concentration
in
eggs
that
caused
50%
effect.
Page
43
of
65
Appendix
3.
Summary
of
Aquatic
Toxicity
Data
for
CDDs
from
the
ACQUIRE
Database
Test
Loc
CAS
#/
Chemical
Latin
name,
Common
name
Endpoint
Effect
Water
Type
Dur
(
days)/

Exp
Typ
Conc
(
ug/
L)
Ref.

LAB
1746016,
2,3,7,8­
Tetrachlorodibenzo[
b,
e][
1,4]
dioxin
Ameiurus
melas,
Black
bullhead
LD50
MOR
FW
80.00/
I
F
5
ug/
kg
Kleeman
et
al.,
1988
Cyprinus
carpio,
Common,
mirror,
colored,
carp
ED50
ENZ
FW
7.00/
I
F
0.048
ug/
kg
Van
der
Weiden
et
al.,

1994a
ED50
ENZ
FW
14.00/
I
F
0.069
mmol/
kg
ED50
ENZ
FW
14.00/
I
F
0.084
mmol/
kg
ED50
ENZ
FW
14.00/
I
F
0.166
mmol/
kg
ED50
ENZ
FW
14.00/
I
F
0.059
mmol/
kg
LD50
MOR
FW
80.00/
I
F
3
ug/
kg
Kleeman
et
al.,
1988
LOEC
ENZ
FW
7.00/
I
F
0.30
ug/
kg
Van
der
Weiden
et
al.,

1994a
LOEC
ENZ
FW
7.00/
I
F
0.03
ug/
kg
LOEC
ENZ
FW
14.00/
I
F
0.03
ug/
kg
LOEC
ENZ
FW
14.00/
I
F
0.03
ug/
kg
LOEC
ENZ
FW
14.00/
I
F
0.03
ug/
kg
LOEC
ENZ
FW
14.00/
I
F
0.03
ug/
kg
NOEC
ENZ
FW
7.00/
I
F
0.01
ug/
kg
Fundulus
heteroclitus,
Mummichog
ED50
PHY
SW
27.00/
S
F
2025
pg/
g
Prince&
Cooper,
1995
Lepomis
macrochirus,
Bluegill
LD50
MOR
FW
80.00/
I
F
16
ug/
kg
Kleeman
et
al.,
1988
Micropterus
salmoides,
Largemouth
bass
LD50
MOR
FW
80.00/
I
F
11
ug/
kg
LAB
1746016,
2,3,7,8­
Tetrachlorodibenzo[
b,
e][
1,4]
dioxin
Oncorhynchus
mykiss,
Rainbow
trout,
donaldson
trout
ED50
BIO
FW
7.00/
I
F
0.17
ug/
kg
Newsted
&
Geiger,
1993
ED50
BIO
FW
7.00/
I
F
0.79
ug/
kg
ED50
CYT
FW
7.00/
I
F
0.5
­
5.0
ug/
kg
ED50
ENZ
FW
10.00/
I
F
1.00
ug/
kg
ED50
ENZ
FW
20.00/
I
F
0.85
ug/
kg
ED50
ENZ
FW
21.00/
I
A
0.79
ug/
kg
Van
der
Weiden
et
al.,

1992
ED50
ENZ
FW
5.00/
I
F
1.05
ug/
kg
Newsted
et
al.,
1995
ED50
ENZ
FW
7.00/
I
F
0.73
ug/
kg
ED50
ENZ
FW
7.00/
I
F
0.66
ug/
kg
ED50
ENZ
FW
3.00/
R
F
3
ug/
kg
Harris
et
al.,
1994
LD50
MOR
FW
30.00/
I
F
242
pg/
g
Zabel
&
Peterson,
1996
LD50
MOR
FW
30.00/
I
F
409
pg/
g
LD50
MOR
FW
~
30.00/
I
A
>
250
­
<
300
pg/
g
LD50
MOR
FW
~
30.00/
I
A
>
200
­
<
300
pg/
g
LD50
MOR
FW
~
30.00/
I
A
>
300
­
<
350
pg/
g
LD50
MOR
FW
~
30.00/
I
A
>
400
­
<
500
pg/
g
LD50
MOR
FW
~
30.00/
I
A
>
300
­
<
400
pg/
g
LD50
MOR
FW
~
30.00/
I
F
0.171
ng/
g
LD50
MOR
FW
~
30.00/
I
F
0.261
ng/
g
Test
Loc
CAS
#/
Chemical
Latin
name,
Common
name
Endpoint
Effect
Water
Type
Dur
(
days)/

Exp
Typ
Conc
(
ug/
L)
Ref.

Page
44
of
65
LD50
MOR
FW
~
30.00/
I
F
0.341
ng/
g
LD50
~
MOR
FW
2.00/
S
F
439
pg/
g
Walker
et
al.,
1992
LD50
MOR
FW
7.00/
I
F
0.240
Walker
&
Peterson.,
1991
LD50
MOR
FW
7.00/
I
F
0.374
LD50
MOR
FW
7.00/
I
F
0.488
LD50
MOR
FW
7.00/
I
F
0.230
LD50
MOR
FW
80.00/
I
F
10
ug/
kg
Spitsbergen,
et
al.,
1988a
LD50
MOR
FW
80.00/
I
F
10
ug/
kg
Kleeman
et
al.,
1988
LOEC
BIO
FW
7.00/
I
F
0.01
ug/
kg
Newsted
and
Giesy,
1993
LOEC
ENZ
FW
10.00/
I
F
0.5
ug/
kg
Newsted
et
al.,
1995
LOEC
ENZ
FW
2.00
­
16.00/
D
A
0.072
ug/
kg
Parrot
et
al.,
1995
LOEC
ENZ
FW
20.00/
I
F
0.1
ug/
kg
Newsted
et
al.,
1995
LOEC
ENZ
FW
5.00/
I
F
0.5
ug/
kg
LOEC
ENZ
FW
7.00/
I
F
0.1
ug/
kg
Newsted
and
Giesy,
1993
NOEC
ENZ
FW
7.00/
I
F
0.01
ug/
kg
NOEC
GRO
FW
56.00/
F
F
<
0.000038
Mehrle
et
al.,
1988
NOEC
MOR
FW
56.00/
F
F
<
0.000038
LAB
1746016,
2,3,7,8­
Tetrachlorodibenzo[
b,
e][
1,4]
dioxin
Oryzias
latipes,
Medaka,
high­
eyes
EC50
ABN
FW
>
13.00/
S
F
0.006
Wisk
&
Cooper,
1990a
EC50
ABN
FW
>
13.00/
S
F
0.015
EC50
ABN
FW
3.00/
S
F
3500
EC50
ABN
FW
3.00/
S
F
14000
EC50
DVP
FW
5.00
­
6.00/
S
F
0.012
Harris
et
al.,
1994
EC50
HAT
FW
3.00/
S
F
14000
Wisk
&
Cooper,
1990b
LC50
MOR
FW
>
13.00/
S
F
0.013
Wisk
&
Cooper,
1990a
LC50
MOR
FW
17.00/
S
F
0.020
Harris
et
al.,
1994
LC50
MOR
FW
0.0
­
17.00/
S
F
0.0057
Metcalfe
et
al.,
1997
LC50
MOR
FW
6.00/
S
F
9000
Wisk
&
Cooper,
1990b
Perca
flavescens,
Yellow
perch
LD50
MOR
FW
80.00/
I
F
3
ug/
kg
Spitsbergen,
et
al.,
1988b
LD50
MOR
FW
80.00/
I
F
3
ug/
kg
Kleeman
et
al.,
1988b
Platichthys
flesus,
Starry,
european
flounder
ED50
ENZ
SW
10.00/
D
F
1.62
ug/
kg
Besselink
et
al.,
1997
Poecilia
reticulata,
Guppy
LT50
~
MOR
FW
11.60/
S
F
1.0
Norris
&
Miller,
1974
LT50
~
MOR
FW
18.20/
S
F
10.0
LT50
~
MOR
FW
21.70/
S
F
0.1
Salvelinus
namaycush,
Lake
trout,
siscowet
LC50
~
MOR
FW
2.00/
R
F
85
pg/
g
Zabel
et
al.,
1995c
LD50
~
MOR
FW
2.00/
R
F
0.065
Walker
et
al,.,
1991
LD50
~
MOR
FW
2.00/
S
F
65
pg/
g
Walker
et
al.,
1992
LD50
MOR
FW
F
65
pg/
g
Walker,
1992
Test
Loc
CAS
#/
Chemical
Latin
name,
Common
name
Endpoint
Effect
Water
Type
Dur
(
days)/

Exp
Typ
Conc
(
ug/
L)
Ref.

Page
45
of
65
LD50
MOR
FW
F
47
pg/
g
LOEC
~
GRO
FW
2.00/
R
F
0.055
Walker
et
al,.,
1991
LOEC
~
GRO
FW
2.00/
R
F
0.226
LOEC
~
MOR
FW
2.00/
R
F
0.055
LOEC
MOR
FW
F
40
pg/
g
Walker,
1992
NOEC
~
MOR
FW
2.00/
R
F
0.034
Walker
et
al,.,
1991
NOEC
MOR
FW
F
34
pg/
g
Walker,
1992
LAB
3268879,
Octachlorodibenzo[
b,
e][
1,4]
dioxin
Brachydanio
rerio,
Zebra
danio,
zebrafish
ET50
HAT
FW
3.48/
F
A
0.033
Berends
et
al.,
1997
ET50
HAT
FW
3.54/
F
A
0.031
LAB
34465468,
Hexachlorodibenzo[
b,
e][
1,4]
dioxin
Cyprinus
carpio,
Common,
mirror,
colored,
carp
ED50
ENZ
FW
14.00/
I
F
0.500
mmol/
kg
Van
der
Weiden
et
al.,

1994b
ED50
ENZ
FW
14.00/
I
F
0.536
mmol/
kg
LOEC
ENZ
FW
14.00/
I
F
0.31
ug/
kg
LOEC
ENZ
FW
14.00/
I
F
1.53
ug/
kg
Oncorhynchus
mykiss,
Rainbow
trout,
donaldson
trout
LD50
MOR
FW
~
30.00/
I
F
8.58
ng/
g
Zabel
et
al,
1995b
LAB
34816530,
1,2,7,8,­
Tetrachlorodibenzo[
b,
e][
1,4]
dioxin
Oryzias
latipes,
Medaka,
high­
eyes
EC50
ABN
FW
>
13.00/
S
F
19.000
Wisk
&
Cooper,
1990a
EC50
ABN
FW
>
13.00/
S
F
>
50.000
LC50
MOR
FW
>
13.00/
S
F
>
50.000
LAB
35822469,
1,2,3,4,6,7,8­
Heptachlorodibenzo[
b,
e][
1,4]
dioxin
Oncorhynchus
mykiss,
Rainbow
trout,
donaldson
trout
LD50
MOR
FW
~
30.00/
I
F
110
ng/
g
Zabel
et
al,
1995b
LOEC
ENZ
FW
2.00
­
16.00/
D
A
0.74
ug/
kg
Parrot
et
al.,
1995
LAB
39001020,
1,2,3,4,5,6,7,8­
Octachlorodibenzofuran
Brachydanio
rerio,
Zebra
danio,
zebrafish
ET50
HAT
FW
3.43/
F
A
0.032
Berends
et
al.,
1997
ET50
HAT
FW
3.54/
F
A
0.036
LAB
40321764,
1,2,3,7,8­
Pentachlorodibenzo[
b,
e][
1,4]
dioxin
Cyprinus
carpio,
Common,
mirror,
colored,
carp
ED50
ENZ
FW
14.00/
I
F
0.025
mmol/
kg
Van
der
Weiden
et
al.,

1994b
ED50
ENZ
FW
14.00/
I
F
0.025
mmol/
kg
LOEC
ENZ
FW
14.00/
I
F
0.002
ug/
kg
LOEC
ENZ
FW
14.00/
I
F
0.01
ug/
kg
Oncorhynchus
mykiss,
Rainbow
trout,
donaldson
trout
LD50
MOR
FW
~
30.00/
I
A
>
450
­
<
500
pg/
g
Zabel,
et
al.,
1995a
LD50
MOR
FW
7.00/
I
F
0.566
Walker
&
Peterson.,
1991
LOEC
ENZ
FW
2.00
­
16.00/
D
A
0.039
ug/
kg
Parrot
et
al.,
1995
Oryzias
latipes,
Medaka,
high­
eyes
EC50
ABD
FW
>
13.00/
S
F
0.013
Wisk
&
Cooper,
1990a
EC50
ABD
FW
>
13.00/
S
F
0.032
LC50
MOR
FW
>
13.00/
S
F
0.027
LAB
57653857,
1,2,3,4,7,8­
Hexachlorodibenzo[
b,
e][
1,4]
dioxin
Oncorhynchus
mykiss,
Rainbow
trout,
donaldson
trout
LD50
MOR
FW
7.00/
I
F
1.427
Walker
&
Peterson.,
1991
LOEC
ENZ
FW
2.00
­
16.00/
D
A
0.25
ug/
kg
Parrot
et
al.,
1995
Oryzias
latipes,
Medaka,
high­
eyes
EC50
ABN
FW
>
13.00/
S
F
0.200
Wisk
&
Cooper,
1990a
EC50
ABN
FW
>
13.00/
S
F
1.100
LC50
MOR
FW
>
13.00/
S
F
2.9
Page
46
of
65
Endpoint:
LC50
Median
lethal
concentration.
Used
only
when
death
is
observed
endpoint.
TLms
and
TL50s
~
LC50
EC50
Median
effective
concentration.
Used
when
an
effect
other
than
death
is
the
observed
endpoint.

NOEC
No
observed
effect
concentration
ED50
Effective
dose
LD50
Lethal
dose
LOEC
Lowest
observed
effect
concentration
LT50
Median
lethal
concentration.
Used
only
when
death
is
observed
endpoint.
TLms
and
TL50s
~
LC50
Effect:
MOR
Mortality.
Effect
expressed
as
percent
death
or
percent
survival.

ENZ
Enzyme
effect.
Change
in
enzyme
activity.

BIO
Change
in
physiochemical
process,
including
glycogen
uptake,
cholesterol
levels,
and
lipid
analysis.

GRO
Growth.
Measurable
change
in
length
and/
or
weight.

ABN
Physical
malformation
during
development
stages.

DVP
Change
in
the
ability
to
grow
to
a
more
mature
life­
stage
and
in
time
between
separate
life­
stages.

HAT
Hatchability.
Change
in
percent
hatch
or
time
to
hatch.

Water
Type:

SW
Saltwater
FW
Freshwater
Exposure
Type:

S
Static
NR
Not
reported
F
Flow
through
I
Injection
R
Renewal
D
Diet
or
oral
exposure
or
both
simultaneously
Source:
ACQUIRE
Database
(
retrieved
October
1999).
Page
47
of
65
Appendix
4.
Aquatic
Toxicity
Data
for
Chlorinated
Dibenzofurans
(
CDFs)
(
U.
S.
EPA,
1990)

Chemical
Form
Test
Species
Duration
(
days)
Concentration
(
ng/
l)
Toxic
Effects
Reference
Exposure
Observation
Monochloro­
No
Data
Dichloro­
No
Data
Trichloro­
No
Data
Tetrachloro­

1,3,7,8­
TCDF
Rainbow
Trout
(
yolk
sac
fry)
4
(
water)
+
10
Days
5.7
x
10­
4
times
as
toxic
as
2,3,7,8­
TCDD
Helder
and
Seinen
(
1986)

2,3,7,8­
TCDF
Rainbow
Trout
(
yolk
sac
fry)
4
(
water)
+
10
Days
9.5
x
10­
5
times
as
toxic
as
2,3,7,8­
TCDD
2,3,7,8­
TCDF
Rainbow
Trout
(
fry
­
0.38
g)
28
(
water)
+
28
Days
8.78
3.93
1.79
0.90
0.41
Controls
28%
dead
(
28
days)

46%
dead
(
56
days)

18%
dead
(
28
days)

22%
dead
(
56
days)

3%
dead
(
28
days)

3%
dead
(
56
days)

6%
dead
(
28
days)

6%
dead
(
56
days)

2%
dead
(
28
days)

2%
dead
(
56
days)

0%
dead
(
56
days)
Mehrle
et
al.
(
1988)

Rainbow
Trout
(
same
effects
as
same
conc.
as
TCDD)
Lethal
edema
Helder
(
1980)

Rainbow
Trout
(
yolk
sac
fry)
4
+
10
Days
8.9
x
10­
2
times
as
toxic
as
2,3,7,8­
TCDD
Helder
and
Seinen
(
1986)

Source:
U.
S.
EPA,
1990.
Page
48
of
65
Appendix
5.
Summary
of
Aquatic
Toxicity
Data
for
CDFs
from
the
ACQUIRE
Database
Test
Loc
CAS
#/
Chemical
Latin
name,
Common
name
Endpoint
Effect
Water
Type
Dur
(
days)/

Exp
Typ
Conc
(
ug/
L)
Ref.

LAB
132649,
Dibenzofuran
Cyprinodon
variegatus,
Sheepshead
minnow
LC50
MOR
SW
1.00/
S
F
>
3200
Heitmuller
et
al.,
1981
LC50
MOR
SW
2.00/
S
F
>
3200
LC50
MOR
SW
3.00/
S
F
3100
LC50
MOR
SW
4.00/
S
F
1800
NOEC
MOR
SW
4.00/
S
F
1000
Daphnia
magna,
Water
flea
LC50
MOR
FW
1.00/
S
F
7500
LeBlanc,
1980
LC50
MOR
FW
2.00/
S
F
1700
LC50
MOR
FW
2.00/
S
F
1340
Maas,
1990
LC50
MOR
FW
2.00/
S
F
12000
Mysidopsis
bahia,
Opossum
shrimp
LC50
MOR
SW
4.00/
NR
F
1310
EPA,
1978
Pimephales
promelas,
Fathead
minnow
LC50
MOR
FW
4.00/
F
F
1780
Geiger
et
al.,
1988
LC50
MOR
FW
4.00/
F
F
1850
LC50
MOR
FW
4.00/
F
F
1050
Brooke,
1991
LC50
MOR
FW
4.00/
S
F
3620
LC50
MOR
FW
4.00/
S
F
1140
LC50
MOR
FW
4.00/
S
F
3020
Poecilia
reticulata,
Guppy
LC50
MOR
FW
4.00/
S
F
1800
Maas,
1990
LC50
MOR
FW
4.00/
S
F
18000
Skeletonema
costatum,
Diatom
EC50
PSE
SW
4.00/
NR
F1500
EPA,
1978
LAB
39001020,
1,2,3,4,5,6,7,8­
Octachlorodibenzofuran
Brachydanio
rerio,
Zebra
danio,
zebrafish
ET50
HAT
FW
3.43/
F
A
0.032
Berends
et
al.,
1997
ET50
HAT
FW
3.54/
F
A
0.036
LAB
51207319,
2,3,7,8­
Tetrachlorodibenzofuran
Cyprinus
carpio,
Common,
mirror,
colored,
carp
ED50
ENZ
FW
14.00/
I
F
17.75
mmol/
kg
Van
der
Weiden
et
al.,

1994b
ED50
ENZ
FW
14.00/
I
F
18.63
mmol/
kg
LOEC
ENZ
FW
14.00/
I
F
10.51
ug/
kg
Oncorhynchus
mykiss,
Rainbow
trout,
donaldson
trout
ED50
ENZ
a
a
F
0.57
umol/
kg
Metcalfe
&
Niimi,
1993
LD50
MOR
FW
7.00/
I
F
8.086
Walker
&
Peterson.,
1991
LOEC
ENZ
FW
2.00
­
16.00/
D
A
0.12
ug/
kg
Parrot
et
al.,
1995
NOEC
GRO
FW
56.00/
F
F
0.00041
Mehrle
et
al.,
1988
NOEC
MOR
FW
56.00/
F
F
0.00179
Oryzias
latipes,
Medaka,
high­
eyes
EC50
ABN
FW
>
13.00/
S
F
0.007
Wisk
&
Cooper,
1990a
EC50
ABN
FW
>
13.00/
S
F
0.019
LC50
MOR
FW
>
13.00/
S
F
0.016
Appendix
5.
Summary
of
Aquatic
Toxicity
Data
for
CDFs
from
the
ACQUIRE
Database
(
continued)

Page
49
of
65
a
Not
reported.

Endpoint:
LC50
Median
lethal
concentration.
Used
only
when
death
is
observed
endpoint.
TLms
and
TL50s
~
LC50
EC50
Median
effective
concentration.
Used
when
an
effect
other
than
death
is
the
observed
endpoint.

NOEC
No
observed
effect
concentration
ED50
Effective
dose
LD50
Lethal
dose
LOEC
Lowest
observed
effect
concentration
Effect:
MOR
Mortality.
Effect
expressed
as
percent
death
or
percent
survival.

ENZ
Enzyme
effect.
Change
in
enzyme
activity.

BIO
Change
in
physiochemical
process,
including
glycogen
uptake,
cholesterol
levels,
and
lipid
analysis.

GRO
Growth.
Measurable
change
in
length
and/
or
weight.

ABN
Physical
malformation
during
development
stages.

Water
Type:

SW
Saltwater
FW
Freshwater
Exposure
Type:

S
Static
NR
Not
reported
F
Flow
through
I
Injection
R
Renewal
D
Diet
or
oral
exposure
or
both
simultaneously
Source:
ACQUIRE
Database
(
retrieved
October
1999).
Page
50
of
65
Appendix
6.
Summary
of
Studies
on
the
Toxicity
of
2,3,7,8­
TCDD
to
Birds
(
U.
S.
EPA,
1990)

Species
Exposure
Level
Symptoms/
Effects
Reference
Chickens
1

g/
kg/
day­­
oral
dose
Positive
chick
edema
(
lesions)
Schwetz
et
al.
(
1973)

Eastern
Bluebird
10,000
ppt
LOAEL
for
eggs
Thiel
et
al.
(
1988)

Chickens
10­
10
M
concentration
of
TCDD
60%
reduction
in
tymphoid
cell
numbers
Nikolaidis
et
al.
(
1988)

White
Leghorn
Chickens
30­
65
ppt
LOAEL
for
eggs
Sullivan
et
al.
(
1987)

Northern
Bobwhite
15

g/
kg
body
weight
single
oral
dose
LD50
Eisler
(
1986)

Ringed
Turtle­
Dove
810

g/
kg
body
weight
single
oral
dose
LD50
Eisler
(
1986)

Mallards
108

g/
kg
body
weight
single
oral
dose
LD50
Eisler
(
1986)

Bobwhite
167
ppt
oral
exposure
in
a
feeding
test
for
5
days
with
total
observation
period
of
8
days
total
LC50
Kenaga
and
Norris
(
1983)

3
ppt
18
wk/
18
wk
test
0.3
ppt
18
wk/
18
wk
test
No
effect
on
reproduction
Mallard
278
ppt
oral
exposure
in
a
feeding
test
for
5
days
with
total
observation
period
of
8
days
LC10
Kenaga
and
Norris
(
1983)

Turkey
>
259
ppt
11
day/
11
day
test
"
EC0"
(
no
effect
level)
Kenaga
and
Norris
(
1983)

Chicken
>
500
ng/
kg/
day
21
day/
21
day
test
>
200
ng/
kg/
day
21
day/
21
day
test
>
100
ng/
kg/
day
21
day/
21
day
test
LC90
Reduced
feeding
and
growth
"
LC0"
(
no
mortality)

"
LC0"
(
no
mortality)
Kenaga
and
Norris
(
1983)

Source:
U.
S.
EPA,
1990.
Page
51
of
65
Appendix
7.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Birds
(
U.
S.
EPA,
1993)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Bobwhite
quail,

Colinus
virginianus
Oral
dose
15,000c
Single
dose
37­
d
LD50
Hudson
et
al.,
1984
Mallard,
Anus
platyrynchos
Oral
dose
>
108,000c
Single
dose
37­
d
LD50
Hudson
et
al.,
1984
Ringed
turtle
dove,

Stretopelia
risoria
Oral
dose
>
810,000c
Single
dose
37­
d
LD50
Hudson
et
al.,
1984
Ring­
necked
pheasant,
Phasianus
colchicus
Eggs
Egg
injection,

yolk
2,100c
Single
dose
28­
d
post
hatch
LD50
Nosek
et
al.,
1992c
10,000c
Single
dose
28­
d
post
hatch
LOAELd
mortality
Nosek
et
al.,
1992c
1,000c
Single
dose
28­
d
post
hatch
NOAELe
Nosek
et
al.,
1992c
Egg
injection,

albumin
1,400c
Single
dose
18­
d
post
hatch
LD50
Nosek
et
al.,
1992c
1,000c
Single
dose
28­
d
post
hatch
LOAEL
mortality
Nosek
et
al.,
1992c
100c
Single
dose
28­
d
post
hatch
NOAEL
Nosek
et
al.,
1992c
Adult
hen
i.
p.
injection
100,000c
Single
dose
77­
d
100%
mortality
after
42­

d
Nosek
et
al.,
1992a
25,000c
Single
dose
77­
d
80%
mortality
Nosek
et
al.,
1992a
6,250c
Single
dose
77­
d
0%
mortality
Nosek
et
al.,
1992a
1,000c
Weekly
injections
for
10
wk
7­
wk
post
dose
LOAEL
hen
mortality
hen
weight
egg
production
embryo
mortality
Nosek
et
al.,
1992a
100c
Weekly
injections
for
10
wk
7­
wk
post
dose
NOAEL
Nosek
et
al.,
1992a
a
Measured
TCDD
concentration
in
water.

b
Measured
TCDD
concentration
in
organism
(
wet
weight).

c
Unmeasured
TCDD
concentration
in
water
or
organism
(
wet
weight).

d
LOAEL=
Lowest
observed
adverse
effect
level.

e
NOAEL
=
No
observed
adverse
effect
level.
Page
52
of
65
Appendix
8.
Summary
of
the
Toxic
Effects
of
2,3,7,8­
TCDD
to
Wildlife
Mammals
(
U.
S.
EPA,
1990)

Test
Species
Test
Method
Water
Conc.

(
ng/
L)
a
Organism
Conc.

(
pg/
g)
b
Duration
Effect
Reference
Exposure
Observation
Mink,
Mustela
vison
Newborn
i.
p.
injection
1,000c
Daily
for
12
d
133­
d
100%
mortality
after
14­

d
Aulerich
et
al.,
1988
1,000c
Daily
for
12
d
133­
d
62%
mortality
after
133­

d
Aulerich
et
al.,
1988
Adult
Oral
dose
4,200c
Single
dose
28­
d
LD50
Hochstein
et
al.,
1988
a
Measured
TCDD
concentration
in
water.

b
Measured
TCDD
concentration
in
organism
(
wet
weight).

c
Unmeasured
TCDD
concentration
in
water
or
organism
(
wet
weight).

Appendix
9.
Summary
of
Studies
on
the
Toxicity
of
2,3,7,8­
TCDD
to
Wildlife
Mammals
(
U.
S.
EPA,
1993)

Species
Exposure
Level
Symptoms/
Effects
Reference
Rhesus
Monkeys
.0017

g/
kg
body
weight
daily
(
7­
29
months)
Abortion
and
weight
loss
Eisler
(
1986)

Monkeys
.011

g/
kg/
day
(
9.3
months)
Generalized
toxicity:
anemia,
hair
loss,

death.
1
of
2
pregnancies
aborted.
Kociba
and
Schwetz
(
1982)

.0017

g/
kg/
day
(
7­
20
months)
Slight
loss
of
weight
and
hair.
4
of
7
pregnancies
aborted.

.24

g/
kg
3
times
weekly
for
3
weeks
of
early
gestation
Severe
toxicity
(
death).
2/
2
pregnancies
aborted.
Kociba
and
Schwetz
(
1982)

.048

g/
kg
3
times
weekly
for
3
weeks
of
early
gestation
Slight
toxicity.
3/
4
pregnancies
aborted.
Species
Exposure
Level
Symptoms/
Effects
Reference
Page
53
of
65
.0095

g/
kg
3
times
weekly
for
3
weeks
of
early
gestation
No
toxicity.
1/
4
pregnancies
aborted.

Control
3/
11
pregnancies
aborted.
Page
54
of
65
Appendix
10.
Summary
of
2,3,7,8­
TCDD
Toxicological
Endpoints
for
Fresh
Water
Fishes
Test
Species
Growth
Stages
Exposure
Toxic
Effect
Toxic.
Endpoints
References
Route
Duration
Observed
LC50
(
ng/
L
in
water)
(
ppt)

Fathead
Minnow
­­
water
28
days
28+
20
days
mor.
1.7
Adams
et
al.
(
1986)

Japanese
Medaka
eggs
­­
fertilized
egg
to
3­
d
post
hatch
­­
mor.
9.0­
13.0
Wisk
and
Copper
1990
a,
b
Rainbow
Trout
Swim­
up
fry
(
0.38
g)
water
(
flow­
through)
28
days
28
days
mor.
0.046
Mehrl
et
al.
(
1988)

Sac
frey
water
(
renewal)
96
hours
21
days
mor.
1.83
Bolet
L.,
(
1989)

EC50
(
ng/
L
in
water
)
(
ppt)

Japanese
Medaka
eggs
­­
fertilized
egg
to
3­
d
post
hatch
­­
embryos
with
lesions
3.5­
6.0
Wisk
and
Copper
1990
a,
b
LOEC
(
ng/
L
in
water)
(
ppt)

Rainbow
Trout
Swim­
up
fry
(
0.38
g)
water
(
flow
through)
28
days
28
days
mor.
0.038
Mehrle
et
al.
(
1988)

NOEC
(
ng/
L
in
water)
(
ppt)

Rainbow
Trout
­­
water
(
flow
through)
56
days
­­
gro/

mor
<
0.038
Mehrle
et
al.
(
1988)

Swim­
up
fry
(
0.38
g)
water
(
flow
through)
28
days
28
days
mor.
0.0011
Mehrle
et
al.
(
1988)

LD50
(
µ
g/
kg
for
dose
and
ng/
g
for
conc.
in
fish)

Blue
Gill
­­
i.
p.
1
dose
80
days
mor.
16
µ
g/
kg
Kleeman
et
al.
(
1988)

Bullhead
­­
i.
p.
1
dose
80
days
mor.
5
µ
g/
kg
Kleeman
et
al.
(
1988)

Carp
­­
i.
p.
1
dose
80
days
mor.
3
µ
g/
kg
Kleeman
et
al.
(
1988)

Large
Mouth
Bass
­­
i.
p.
1
dose
80
days
mor.
11
µ
g/
kg
Kleeman
et
al.
(
1988)

Rainbow
Trout
­­
i.
p.
1
dose
80
days
mor.
10
µ
g/
kg
Spitsbergen
et
al.
(
1988a)

­­
i.
p.
1
dose
80
days
mor.
3
µ
g/
kg
Kleeman
et
al.
(
1988)

Yellow
Perch
­­
i.
p.
1
dose
80
days
mor.
10
µ
g/
kg
Kleeman
et
al.
(
1988b
Yellow
Perch
­­
i.
p.
1
dose
80
days
mor.
3
µ
g/
kg
Spitsbergen
et
al.
(
1988b)
Appendix
10.
Summary
of
2,3,7,8­
TCDD
Toxicological
Endpoints
for
Fresh
Water
Fishes
(
continued)

Test
Species
Growth
Stages
Exposure
Toxic
Effect
Toxic.
Endpoints
References
Route
Duration
Observed
Page
55
of
65
Rainbow
Trout
­­
injec.
1
dose
30
days
mor.
0.171
ng/
g
(
in
fish)
Zabel
&
Peterson
(
1996)

­­
injec.
1
dose
7
days
mor.
0.230
ng/
g
(
in
fish)
Walter
&
Peterson
(
1991)

Finger­
ling
i.
p.
1
dose
80
days
mor.
10
ng/
g
(
in
fish)
Spitsbergen
et
al.
1988a
Kleeman
et
al.
(
1988)

Eggs
water
(
renewal)
48
hours
>
48
hrs
to
post
swimup
mor.

(
sac
fry)
0.439
ng/
g
(
in
eggs)
Walker
et
al.
(
1992)

ED50
(
µ
g/
kg
for
dose
and
ng/
g
for
conc.
in
fish)

Carp
­­
injec.
1
dose
7
days
enz
0.048
µ
g/
kg
Van
der
Weiden
et
al.
1994a
Rainbow
Trout
­­
injec.
1
dose
7
days
bio
0.17
µ
g/
kg
Newsted
&
Geiger,
1993
­­
injec.
1
dose
7
days
enz
0.66
µ
g/
kg
Newsted
et
al.
1995
Juvenile
(
46
g)
i.
p.
1
dose
3
weeks
EROD
Induction
0.79
ng/
g
(
in
fish)
Van
der
Wekdeb
et
al.
1992
immature
adult
(
350
g)
i.
p.
1
dose
72­
h
AHH
Induction
0.64
ng/
g
(
in
fish)
Janz
and
Metcalfe,
1991
LOAEL
(
µ
g/
kg
in
fish)

Carp
­­
injec.
1
dose
7
days
enz
0.3
µ
g/
kg
(
in
fish)
Van
der
Weiden
et
al.
1994a
­­
injec.
1
dose
14
days
enz
0.03
µ
g/
k
(
in
fish)
Van
der
Weiden
et
al.
1994a
Rainbow
Trout
­­
injec.
1
dose
7
days
enz
0.01
µ
g/
kg
(
in
fish)
Newsted
and
Giesy,
1993
­­
injec.
1
dose
14
days
enz
0.1
µ
g/
kg
(
in
fish)
Newsted
et
al.,
1995
­­
diet
2­
16
days
­­
enz
0.072
µ
g/
kg
(
in
fish)
Parrot
et
al.,
1995
Appendix
10.
Summary
of
2,3,7,8­
TCDD
Toxicological
Endpoints
for
Fresh
Water
Fishes
(
continued)

Test
Species
Growth
Stages
Exposure
Toxic
Effect
Toxic.
Endpoints
References
Route
Duration
Observed
Page
56
of
65
Lake
Trout
eggs
water
(
renewal
48­
hours
>
48
hours
to
post
swimup
mor.

(
Sac
fry)
0.055
ng/
g
(
in
egg)
Walker
et
al.,
1991
NOAEL
(
µ
g/
kg
in
fish)

Carp
­­
injec.
1
dose
7
days
enz
0.01
µ
g/
kg
(
in
fish)
Van
der
Weiden
et
al.
1994a
Rainbow
Trout
­­
injec.
1
dose
7
days
enz
0.01
µ
g/
kg
(
in
fish)
Newsted
and
Giesy,
1993
Lake
Trout
eggs
water
renewal
48­
hours
>
48
hours
to
post
swimup
­­
0.034
ng/
g
(
in
egg)
Walker
et
al.,
1991
Exposure
Routes
(
Test
Methods):

i.
p.
Intraperitoneal
injection
injec.
injection
water
water
water
(
static)
static
water
water
(
flow­
through)
flow­
through
water
water
(
renewal)
water
renewal
diet
fish
diet
Toxic
Effects:

mor.
mortality
enz
effects
on
Enzyme
Activity
bio
biological
effects
gro
growth
retardation
phy
physical
deformation
Page
57
of
65
Appendix
11.
Summary
of
2,3,7,8
TCDD
Toxicological
Endpoints
for
Marine/
Estuarine
Fish
Test
Species
Growth
Stages
Exposure
Toxic
Effect
Toxic.
Endpoints
References
Route
Duratio
n
Observed
ED50
European
flounder
­­
diet
10
days
­­
enz
1.62
µ
g/
kg
Besselink
et
al.
(
1997)

Mummichog
­­
water
(
static)
27
days
­­
phy
2.025
ng/
g
Prince
and
Cooper
(
1995)

Exposure
Routes
(
Test
Methods):
water
(
static)
static
water
diet
fish
diet
Toxic
Effects:
enz
effects
on
enzyme
activity
phy
physical
deformation
Page
58
of
65
Appendix
12.
Summary
of
2,3,7,8
TCDD
Toxicity
Data
for
Fresh
Water
Invertebrates
Test
Species
Growth
Stages
Exposure
Toxic
Effect
Test
Conc.
(
ng/
L)
References
Route
Duration
Observed
(
Not
a
NOEC/
LOEC
value)

Annelid,
Worm
­­
water
(
static)
55
days
­­
no
effect
on
repro.
success
200
ng/
L
(
water)
Miller
et
al.
(
1973)

Mollusc,
Snail
adult
Model
Ecosy.
33
days
­­
No
toxic
effect
1,330
ng/
L
(
water)
502
ng/
g
(
organism)
Isensee
and
Jones
(
1975)

adult
water
(
static)
36
days
12
days
no
effect
on
repro.
success
200
ng/
L
(
water)
Miller
et
al.
(
1973)

Arthropod,
Mosquito
­­
water
(
static)
17
days
23
days
no
effect
200
ng/
L
(
water)
Miller
et
al.
(
1973)

Arthropod,
Cladoceran
­­
Model
Ecosy.
33
days
­­
no
effect
1,330
ng/
L
(
water)
1,570
ng/
g
(
organism)
Isensee
and
Jones
(
1975)

Arthropod,
Cladoceran
1­
21
days
water
(
renewal)
48
hours
7
days
no
effect
1,030
ng/
L
(
water)
Miller
et
al.
(
1979)
Page
59
of
65
Appendix
13.
Summary
of
Acute
and
Short
Term
2,3,7,8­
TCDDToxicological
Endpoints
for
Birds
Exposure
LD50
(
ng/
kg)
Reference
Route
Duration
Observation
Northern
bobwhite
quail
oral
single
dose
37
days
15
(
95%
confidence
limits:
9.2­
24.5)
Hudson
et
al.
(
1984)

Bobwhite
oral
5
days
7
days
LC50
=
0.167
(
in
feed)
Kenaga
and
Norris
(
1983)

Ringed
turtle
dove
oral
single
dose
37
days
>
810
Hudson
et
al.
(
1984)

Mallard
oral
single
dose
37
days
>
108
Hudson
et
al.
(
1984)

Ring­
necked
pheasant
i.
p.
injection
single
dose
11
weeks
25
(
LD75)
Nosek
et
al.
(
1992)

Appendix
14.
Summary
of
Subchronic
and
Chronic
2,3,7,8­
TCDD
Toxicological
Endpoints
for
Birds
Test
Species
Exposure
LOAEL
(
µ
g/
kg­
day)
NOAEL
(
µ
g/
kg­
day)
Toxic
Effects
References
Route
Duration
Observed
Domestic
Chicken
oral
daily
doses
for
20
to
21
days
20
­
21
days
1.0
0.10
mortality
Schwetz
et
al.
(
1973)

Pheasant
i.
p.
injec.
10
weeks
­­
0.14
0.014
fertility,
embryo
mortality
Nosek
et
al.
(
1992b)
and(
1993
Page
60
of
65
Appendix
15.
Summary
of
Acute
and
Short
Term
2,3,7,8­
TCDD
Toxicological
Endpoints
for
Mammals
Growth
Stages
Exposure
LD50
(
µ
g/
kg)
Reference
Route
Duration
Observation
Dog
­­
oral
single
dose
­­
100
­
200
Kociba
and
Schwetz
(
1982)

Guinea
pig
­­
oral
single
dose
­­
0.6
­
2.1
Schwetz
et
al.
(
1973)

Hamster
­­
oral
single
dose
­­
1,160
­
5,050
Kociba
and
Schwetz
(
1982)

Mink
male
adults
oral
single
dose
28
days
4.2
Hochstein
et
al.
(
1988)

kits
injection
single
dose
12
days
<
0.1
Aulerich
et
al.
(
1988)

Mouse
­­
oral
single
dose
­­
114
­
284
Kociba
and
Schwetz
(
1982)

Rabbit
­­
oral
single
dose
­­
115
Schwetz
et
al.
(
1973)

Rat
­­
oral
single
dose
­­
22
­
45
Schwetz
et
al.
(
1973)

Rhesus
monkey
­­
oral
single
dose
­­
70
Kociba
and
Schwetz
(
1982)

Appendix
16.
Summary
of
Subchronic
and
Chronic
2,3,7,8­
TCDDToxicological
Endpoints
for
Mammals
Exposure
Duration
Toxic
Effects
LOAEL
(
µ
g/
kg­
day)
NOAEL
(
µ
g/
kg­
day)
References
Rat
2
years
female
mortality
0.1
0.01
Kociba
et
al.
(
1978)

Rat
gestation
days
6
to
15
litter
size
pup
weight
0.25
0.125
Khera
and
Ruddick
(
1973)

Rat
3
generations
reproductive
capacity
0.01
0.001
Murray
et
al.
(
1979)

Rhesus
Monkey
7
months
number
of
births
0.0021
­­
Allen
et
al.
(
1979)

Rhesus
Monkey
7
­
48
months
maternal
reproductive
0.00059
0.00012
Bowman
et
al.
(
1989)
Page
61
of
65
APPENDIX
17:
Terrestrial
Exposure
Assessment
The
meadow
vole
(
Microtus
pennsylvanicus)
was
selected
as
the
mammalian
receptor
species
for
evaluating
potential
effects
of
CDD/
CDFs
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
CDD/
CDFs
to
birds.
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
vole
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,
1993b)
and
0.093
kg/
kg
bwd
for
adult
bobwhites
(
U.
S.
EPA,
1993b)
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
(
1993b)
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
CDD/
CDFs.
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.
Page
62
of
65
The
concentration
of
a
chemical
in
a
terrestrial
plant
(
C
p)
as
fresh
weight
was
determined
using
the
following
equation:

C
p
=
C
soil
*
Kps
(
3)

where:

C
soil
=
concentration
ofcontaminant
of
concern
detected
in
surface
soil
(
mg/
kg);
and
Kps
=
partition
coefficient
in
plant
tissue
relative
to
contaminant
soil
concentration
of
TCDD
(
unitless)
(
Chiao
et
al.,
1994).

The
predicted
EEC
for
CDD/
CDF
total­
TEQ
in
soil
(
see
Environmental
Fate
Modeling
chapter
of
this
document)
was
used
as
the
C
soil
in
the
model
for
voles
and
bobwhites.

Employing
equations
1,
2,
and
3,
the
estimated
dose
that
voles
and
bobwhites
would
receive
from
ingestion
of
plants
in
the
vicinity
of
pentachlorophenol­
treated
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
CDD/
CDFs
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
=
+

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.
Page
63
of
65
In
the
risk
assessment,
the
total
dietary
intakes
are
compared
to
toxicity
endpoints
(
e.
g.,
LD50,
NOAEL)
to
determine
if
adverse
effects
are
likely
to
occur
in
voles
and
bobwhites
from
the
ingestion
of
CDDs/
CDFs
in
terrestrial
plants
and
surface
soil.
The
toxicity
endpoints
used
were
obtained
from
laboratory
toxicity
studies
conducted
using
TCDD..
These
endpoints
were
not
converted
to
TEQ.
Toxicity
reference
values
from
studies
based
on
avian
species
other
than
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,
1993b).

Following
adjustment
of
the
mammalian
toxicity
reference
values,
the
total
dose
for
the
bobwhite
and
vole
of
CDD/
CDF
TEQ
was
compared
with
the
toxicity
reference
values
for
TCDD
to
determine
the
risk
quotients
(
RQs).
The
RQs
for
the
bobwhite
and
vole
based
on
exposure
to
CDD/
CDF
TEQ
are
shown
in
Tables
A1
and
A2,
respectively.
Page
64
of
65
Table
A1.
Acute
and
Chronic
Risk
Quotient
Calculations
for
Birds
Compound
EEC
(
mg/
kg)
Bobwhite
Soil
Ingestion
Rate
(
kg/
kg­
d)
Bobwhite
Soil
Dose
(
mg/
kg­
day)
Plant/
soil
partifion
coefficitent
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
CDD/
CDF
TEQ
2.32e­
02
0.0093
2.16e­
04
3.40e­
01
7.89e­
03
1
7.89e­
03
0.093
7.34e­
04
9.49e­
04
LD50
1.5e­
05
mg/
kg
(
Hudson
et
al.,

1984)
acute
63.321
NOAEL
1.4e­
05
mg/
kg­
d
(
Scwetz
et
al.,

1973)
chronic
67.782
TCDD
2.38e­
04
0.0093
2.20e­
06
3.40e­
01
8.09e­
05
1
8.09e­
05
0.093
7.50e­
06
9.70e­
06
LD50
1.5e­
05
mg/
kg
(
Hudson
et
al.,

1984)
acute
0.653
NOAEL
1.4e­
05
mg/
kg­
d
(
Scwetz
et
al.,

1973)
chronic
0.69
Page
65
of
65
1.
Exceeds
acute
High
Risk
Level
of
Concern
(
LOC)

2.
Exceeds
Chronic
Risk
LOC
3.
Exceeds
acute
Restricted
Use
LOC
Table
A2.
Acute
and
Chronic
Risk
Quotient
Calculations
for
Mammals
Compound
EEC
(
mg/
kg)
Vole
Soil
Ingestion
Rate
(
kg/
kg­
d)
Vole
Soil
Dose
(
mg/
kg­
day)
Plant/
soil
partition
coefficient
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)
Endpoint
(
mg/
kg)
RQ
CDD/
CDF
TEQ
2.32e­
02
0.035
8.12e­
04
3.40e­
01
7.89e­
03
1
7.89e­
03
0.35
2.76e­
03
3.57e­
03
LD50*

0.09
mg/
kg
(
Kociba
and
Schwetz,
1982)
acute
0.04
NOAEL*

9.94e­
04
mg.
kg­
day
(
Khera
and
Ruddick,
1973)
chronic
3.592
TCDD
2.38e­
04
0.035
8.30e­
06
3.40e­
01
8.09e­
05
1
8.09e­
05
0.35
2.83e­
05
3.66e­
05
LD50*

0.114
mg/
kg
(
Kociba
and
Scwetz,
1982)
acute
4.07e­
04
NOAEL*

9.94e­
04
mg.
kg­
day
(
Khera
and
Ruddick,
1973
chronic
0.04
*
Value
for
mouse
(
LD50
of
0.114
mg/
kg­
day)
and
rat
(
NOAEL
of
0.125
mg/
kg­
day)
were
adjusted
for
body
weight
as
described
in
Sample
et
al.(
1996)
and
U.
S.

EPA
(
1993b)