5457
905R84003
J. VINCENT NABHGLZ
TS-796, US EPA
401 M ST, SW
WASHINGTON. DC 20460
ENVIRONMENTAL RISK AND HAZARD ASSESSMENTS
FOR VARIOUS ISOMERS OF
POLYCHLORINATED BIPHENYLS (MONQCHLOROBIP&ENYL .THROUGH
HEXACHLOROBlPriENYL AND DECACHLOROBIPHENYL)
by
ENVIRONMENTAL EFFECTS BRANCH
HEALTH AND ENVIRONMENTAL REVIEVV DIVISION
OFFICE OF TOXIC SUBSTANCES
APRIL, 1964
OFFICE OF PESTICIDES AND ,-TO-iIC SUBSTANCES
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON*; D»C. 20460
fa| Protection Agency -t
Ksgion V, Library •
230 SoMfi D? if torn Street
Chicago, Illinois 60604
-------
» (
U,S. EfivfwiifrYttjrrtBf Pigfcaetfori Agency
-------
I. Executive Summary
Potential environmental risk is expected to occur to aquatic
organisms if polychlorinated biphenyl (PCB) isomers are released
into the aquatic environment at concentrations of 100 ug/L.
Expected risks include: (1) reduction in the reproductive
success of fish through direct toxicity to eggs and embryos which
are fish early life stages; (2) direct lethality and sublethal
effects to juvenile and adult fish; (3) effects to aquatic
invertebrates which are food sources for fish; and (4) possible
but not demonstrated reductions in the growth of phytoplankton,
i.e., minute floating plants, usually algae, which produce the
basic food for the aquatic ecosystem. Risks to wild mammals and
birds could not be determined because of a lack of toxicological
information for PCB isomers and lack of an appropriate exposure
assessment for terresrial organisms.
Environmental risk for all effects generally increase with an
increase in the degree of chlorination of the biphenyl (Figure
1). Environmental risk to aquatic organisms from
monochlorobiphenyl isomers are predicted to occur at 10 to 30% of
chemical plant sites expected to inadvertently produce PCB
isomers as impurities; while risk from hexachlorobiphenyl isomers
are expected to occur at 80-90% of plant sites. This increase in
risk is due to an increase in the toxicity of PCBs with greater
numbers of chlorine atoms. An increase in chlorination of PCBs
has been related to increases in toxicity for (1) chronic
toxicity to rainbow trout (Figure 3) for monochlorobiphenyl
through hexachlorobiphenyl isomers, (2) acute (lethal) toxicity
to rainbow trout (Figure 2) for monochlorobiphenyl through
tetrachlorobiphenyl isomers, and (3) acute toxicity to aquatic
invertebrates (Figure 4) for monochlorobiphenyl through
tetrachlorobiphenyl isomers. Acute toxicity of
pentachlorobiphenyl and hexachlorobiphenyl isomers to aquatic
invertebrates is less than would be predicted (Figure 4) because
of the greater water insolubility of these isomers.
-------
Environmental risk to aquatic organisms is shown to increase
to as many as 40% of additional chemical plant sites in the US
under low streamflow conditions. Risk was estimated for two
streamflow conditions for rivers in the US: average streamflow
and low streamflow (Table 1). The increase in risk under low
streamflow conditions can be attributed to lower stream dilution
factors during low streamflow which result in higher predicted
environmental concentrations.
The environmental risk of monochlorobiphenyl and
dichlorobiphenyl isomers is predicted to increase if discounting
factors are incorporated into the exposure assessment. A
discounting factor of 50 for monochlorobiphenyl isomers is
expected to increase potential risk to an additional 40% of plant
sites over risks associated with no discounting factor. A
discounting factor of 5 for dichlorobiphenyl isomers is expected
to increase risk to aquatic organisms at about 10% additional
plant sites. A discounting factor means that monitored
concentrations of a PCB isomer will be divided by that factor,
e.g., 50 or 5, and then recorded. For example, an effluent
concentration for a monochlorobiphenyl of 5 mg/L would be
reported as 100 ug/L of isomer. The increase in risk associated
with the use of discounting factors can be entirely attributed to
higher surface water concentrations used in the exposure
assessment.
The environmental concerns for PCB isomers generated by tnis
risk assessment are similar to the environmental concerns
historically expressed for the commercial mixtures of PCBs
(Aroclors). Environmental concerns held in common are: (1)
impairment of the reproductive success in fish, especially, to
the early life stages of fish; (2) direct adverse effects to
juvenile and adult fish; and (3) reduced survival and growth of
aquatic invertebrates and aquatic plants. Environmental concerns
attributed to PCB commercial mixtures but not yet demonstrated
for PCB isomers include: (1) correlation of high body burdens of
-------
PCBs in female fish with failure of eggs to hatch; (2) impaired
bone development and testical abnormalities in juvenile fish; (3)
contamination of economically important food resources, e.g.,
closure of fisheries; and (4) impairment of reproductive success
in some wild mammals (e.g., mink) and birds.
This environmental risk assessment is a qualitative risk
assessment and is relatively conservative, i.e., designed to err
on the side of environmental protection. The basis of this risk
assessment was the application of information derived from a
hazard assessment to situations identified by an exposure
assessment. Concentration-effect curves for each group of
aquatic organisms, e.g., early life stages of fish, were
synthesized for each class of PCB isomer, e.g.,
monochlorobiphenyls, dichlorobiphenyls, etc., and were compared
to hypothetical surface water concentrations downstream from
organic chemical plants. Whenever an effective concentration was
equal to or exceeded a surface water concentration, a potential
risk was noted and the type and degree of risk was determined
(Tables 1 and 2). No attempt was made to quantify the impact to
a particular population at a particular site over time.
Therefore, this risk assessment can be characterized as more
qualitative than quantitative.
This assessment can also be characterized as relatively
conservative because (1) toxicity information for the most
sensitive species and the most sensitive life stage for a species
was used whenever possible; and (2) it was assumed that all
chemical plants discharged process wastewater containing 100 ug/L
of PCBs and that there was no loss due to sorption,
transformation, or degradation. However, the risk assessment
could have been more conservative. No assessment factor was used
to predict a "safe" concentration of PCB isomer from the toxicity
information. An assessment factor is defined as a number by
which an effect concentration, e.g., EC50 or NOEC, is adjusted
(by division) to arrive at a "safe" environmental
-------
concentration. In addition, information for the most sensitive
species could not always be used. For example, in some data sets
rainbow trout was the most sensitive fish species tested,
however, data from five other fish species were also used in the
risk assessment.
This risk assessment is as comprehensive as available
information for the toxicity of PCB isomers would permit.
Concentration-effect curves were synthesized from all available
information: which ranged from chronic no-observed-effect
concentrations for the most sensitive life stage to acute
concentrations which caused 100% lethality in a test
population. For example, potential environment risk to fish
populations to tetrachlorobiphenyl isomers with no chlorines in
the _o,_o'- positions on the biphenyl (Figure 1 and Table 1) was
characterized in terms of effects to fish early life stages (E),
sublethal effects to juvenile fish (SL), reductions in the growth
of juvenile fish (G), and lethality to juvenile fish (L).
All 18 hypothetical exposure situations (i.e., streamwater
concentrations for the 10th through the 90th percentiles of
chemical plants under two streamwater condition: low and average
streamflow, Table 1) were evaluated for potential risk from each
class of PCB isoomer. In all, one hundred forty four situations
were evaluated for four groups of aquatic organisms. Including
evaluation of discounting factors, a total of 612 situations were
evaluated for potential environmental risks.
This risk assessment did not address the additional risk of
PCB isomers from bioconcentration, food chain transport, and food
chain concentration (or biomagnification). For example, a female
fish will bioconcentrate PCBs and subsequently pass these
residues to her eggs which could result in inviable eggs and
embyros. The reason for this deficiency in the risk assessment
was a lack of toxicological information on the PCB isomers with
regard to this type of reproductive inhibition in fish, birds,
and wild mammals.
-------
II. Environmental Risk Assessment of PCB Isomers to Aquatic
Organisms
A. Basis of Risk Assessment
Risk assessment is the application of information
derived from a hazard assessment to situations identified by an
exposure assessment. The hazard assessment for various PCB
isomers can be found in Sections III through VI. The exposure
assessment was derived from an exposure assessment for
incidentally produced PCBs (Versar 1983). Table H-l from Versar
(1983) (see Table 1) was used to obtain hypothetical surface
water concentrations downstream from organic chemical plants in
the US expected to inadvertently produce PCB isomers as
impurities. Versar estimated 19 hypothetical concentrations: 9
under average streamflow conditions and 9 under low streamflow
conditions. Versar estimated the 10th through the 90th
percentiles (9 percentiles) for each streamflow condition (Table
1). For this risk assessment, the 18 hypothetical streamwater
concentrations were rank-ordered from highest to lowest: 66 to
0.00037 ug/L, respectively. Versar assumed that all plants
discharged process wastewater containing 100 ug/L of PCBs, that
there was instantaneous mixing, and that there was no loss due to
degradation, transformation, or sorption. For this risk
assessment, eight risk assessments were performed: one each for
monochloro-, dichloro-, trichloro-, tetrachloro(with 1-4
chlorines at the _o_,^'-positions)-, tetrachloro(with no chlorines
at the _o,_o_'-pos it ions ) - , pentachloro-, hexachlorobiphenyl
isomers, and decachlorobiphenyl (Table 1). It was assumed that
all plants discharged 100 ug/L of one class of PCB isomers, e.g.,
monochlorobiphenyl isomers. Each estimated streamwater
concentration was compared to the concentration-effect curve for
each class of PCB isomers, and if a risk was predicted (i.e., if
an effective concentration was equal to or greater than a
streamwater concentration), then the type of risk and its degree
were indicated in Table 1. An empty cell in the matrix under
"Class of PCB" in Table 1 indicates that the streamwater
-------
concentration was less than any available effective concentration
in the hazard assessment and thus no risk.
B. Risk Assessment for Fish
1. Monochlorobiphenyls
a. Discharge of 100 ug/L of monochlorobiphenyl
isomers will result in risk to fish populations under only low
streamwater conditions at 30 percent of plants with a low
streamwater dilution factor of 18.9 or less (Table 1). At a
concentration of 66 ug/L, juvenile and adult fish are predicted
to have only sublethal effects, e.g., reduction in growth,
reduced food consumption, disorientation. No lethality of
juveniles and adult fish is expected. The early life stages of
fish, i.e., embryos and sac fry, are expected to be affected by
the monochlorobiphenyl isomers at all concentrations equal to and
higher than 5.3 ug/L. Effects expected to occur are reduced
growth and survival of embryos and sac fry. The percentage of
fish affected cannot be estimated, but at 5.3 ug/L the percentage
will approach 0%.
b. Discharge of monochlorobiphenyl isomers with a
discounting factor of 50 will increase the risk to fish
populations 50 times. A discount of 50 means that monitored
concentrations of monochlorobiphenyls will be divided by 50 and
then reported. Actual concentrations would be 50 times higher
than PCB surface water concentration estimated in Table 1. Under
these conditions, acute lethality of juvenile and adult fish
would occur at 3U% of plants at low streamflow. Lethality of 50%
or higher, would occur at 20% of these plants. Sublethal effects
and subchronic mortality would occur at 40% of plants at low
streamflow and 10% of plants at average streamflow. Effects to
fish early life stages would occur at 70% of plants at low flow
and 40% of plants at average streamflow.
-------
2. Dichlorobiphenyls
a. Discharge of dichlorobiphenyls isomers at 100
ug/L will potentially affect fish populations at 20% of plants at
average streamflow conditions and 50% of plants at low streamflow
(Table 1). Fish early-life stages will be at risk at all the
above plants. Effects will be near zero at 0.47 ug/L and
increase in severity at plants with smaller stream dilution
factors (i.e., as you proceed up Table 1). Juvenile and adult
fish will suffer sublethal effects at only 20% of plants under
low streamflow. The probability of subchronic and acute
lethality is low for dichlorobiphenyls released at 100 ug/L.
b. If a discounting factor of 5 is used for
discharge of dichlorobiphenyl isomers, potential risk to fish
will extend to 10% more plants and acute lethality will probably
occur. Effects to fish early-life stages will occur at 30% of
plants during average streamflow and 60% of plants during low
flow. Sublethal effects to juveniles and adults will occur at
30% of plants only at low flow and lethality approaching
approximately 40% will occur at 10% of plants.
3. Trichlorobiphenyls
Trichlorobiphenyl isomers discharged at 100 ug/L
have a potential to affect the early-life stages of fish at 40%
of plants during average streamflow and 60% plants at low flow
(Table 1). Sublethal and subchronic (30d) lethality will occur
only during low streamflow conditions at 30% of plants.
Subchronic lethality will occur only at 10% of these plants at
low flow and will be much lower than 50%.
-------
4. Tetrachlorobiphenyl Isomers with 1-4 Chlorines at
the OyO'-Positions of the Biphenyl
Tetrachlorobiphenyl isomers with 1 to 4 chlorine
atoms at the o_,^'-positions of the biphenyl (Figure 1) have the
potential to affect fish populations at 70% of plants during low
streamwater flows and at 40% of plants during average flow. The
early-life stages (i.e., embryos and sac fry) of fish will
exhibit reduced growth and increased mortality at all of the
above plant sites. Juvenile and adult fish will exhibit
sublethal effects at only 10% of plant sites during average •
streamflow, but during low streamflow 40% of plants discharging
100 ug/L could affect juveniles and adults (Table 1). Subchronic
lethality will occur only at low streamflow at 20% of plant
sites; these isomers have the potential of killing over 50% of
the fish populations at 10% of plant sites (Table 1).
5. Tetrachlorobiphenyl Isomers with No Chlorines at
. the 0,0'- Positions of the Biphenyl
Tetrachlorobiphenyl isomers with no chlorines at
the .Oyjo1- positions of the biphenyl appear to present
significantly greater risk to fish populations than
tetrachlorobiphenyl isomers with chlorines at the o_ro_'~
positions. They also have the potential of affecting more plant
sites (20% more sites at average streamflow). The early life
stages of fish will be affected at 70% of plant sites at average
streamflow and 80% of plant sites at low flow (Table 1).
Sublethal effects to juvenile and adult fish will (1) begin to
occur at 40% of plants during average streamflow, (2) reduce the
growth of juveniles by 30% at 30% of plant sites, (3) reduce
growth further by 60% at 20% of plants during average flows, and
(4) could kill 25% of the fish populations at 10% of plant sites
during average flow, but during low flows 40-50% of sites could
be affected (Table 1).
-------
6. Pentachlorobiphenyls
Pentachlorobiphenyl isomers were predicted to be
slightly more chronically toxic than the tetrachlorobiphenyl
isomers in the hazard assessment (Section III. E) and, therefore,
will present slightly more risk to resident fish populations than
the tetrachlorobiphenyls given the same discharge conditions at
100 ug/L concentration. Table 1 reflects this increase in risk
by affecting about 10% more plant sites than the
tetrachlorobiphenyls). The subchronic and chronic effects of the
pentachlorobiphenyls are less quantified than they were for the
tetrachlorobiphenyl isomers, because much more experimentation
was available for the tetrachlorobiphenyls and only chronic NOECs
could be predicted for pentachlorobiphenyls in the hazard
assessment. The type of effects could not be quantified but it
was assumed that effects observed for the tetrachlorbiphenyl
isomers will also occur with the pentachlorobiphenyl isomers.
7• Hexachlorobiphenyls
Hexachlorobiphenyl isomers were predicted to be
more chronically toxic than the pentachlorobiphenyls in the
hazard assessment (Section III.F) and, therefore, are assumed to
present more risk to fish populations given the same exposure
conditions. Table 1 indicates this increase in toxicity, and,
therefore, risk.
The hexachlorobiphenyls are expected to affect the
early life stages of fish at 80% of the plant sites during
average streamflow and over 90% of sites at low flow (Table 1).
These effects will increase in severity at sites with smaller
stream dilution factors. This fact is demonstrated at the 20th
percentile entry. At this surface water concentration (16 ug/L),
100% of the embryo and sac fry fish are expected to be killed.
Broyles and Noveck (1979b_) observed 100% mortality within 79 days
10
-------
of lake trout and chinook salmon sac fry after only an 8 ug/L
exposure for 15 days (Section III.F.2.c).
8. Decachlorobiphenyl
Decachlorobiphenyl is expected to be similar to or
less toxic than the hexachlorobiphenyl isomers (Section III.G),
and, therefore, under the same exposure conditions, are expected
to present similar or smaller risks to fish populations. Table 1
reflects the assumption that the potential risks for
decachlorobiphenyl are similar to those for the hexachloro-
biphenyl isomers. No toxicological information is available
indicating how much less toxic decachlorobiphenyl is to fish than
the hexachlorobiphenyl isomers.
C. Risk Assessment for Aquatic Invertebrates
1. Discharges of 100 ug/L
Discharges of 100 ug/L of the various PCB isomers
will probably result in risks to aqutic invertebrates. These
risks are similar to risks predicted for juvenile and adult
fish. Risks to aquatic invertebrates can best be defined by
taking Table 1 and eliminating risks to the early life stages of
fish (i.e., eliminate Es' from Table 1; see Table 2). The
rationale for using the risk assessment for juvenile fish for
aquatic invertebrates is based upon the similar acute toxicity
between fish and aquatic invertebrates (Section IV.A) and a lack
of chronic toxicity information for aquatic invertebrates
(Section IV.B).
There are three major differences between the risk
assessment for fish (Table 1) and the risk assessment for aquatic
invertebrates (Table 2): (1) risk from acute exposure is expected
to occur to aquatic invertebrates (acute risk was not predicted
for fish), (2) chronic sub lethal effects for aquatic
11
-------
invertebrates cannot be quantified as precisely as was done for
fish, and (3) risk to aquatic invertebrates from the
pentachlorobiphenyl isomers, hexachlorobiphenyl isomers, and
decachlorobiphenyl may not occur at the lower streamwater
concentrations.
a. Risk from acute exposure is expected to occur
to aquatic invertebrates from trichlorobiphenyl and
tetrachlorobiphenyl isomers at 10% and 20% of plant sites,
respectively, during low streamflow (Table 2). During low
streamflow, a 100 ug/L discharge of trichlorobiphenyl isomers
will kill 50% of the aquatic invertebrates in the receiving
stream; discharge of tetrachlorobiphenyl isomers will kill over
50% of the invertebrate population. Acute lethality will also
occur at 20% of plants during low streamflow if tetrachlorobi-
phenyl isomers are released (Table 2). In the risk assessment for
fish, lethality was expected to occur only from subchronic exposures
b. Sublethal effects for aquatic invertebrates
cannot be identified or quantified because no chronic studies
have been done. Sublethal effects will probably include
reductions in weight, fertility, brood size, growth rate, and
survival of offspring. It was assumed that the chronic NOECs for
aquatic invertebrates will be similar to the subchronic NOECs for
juvenile fish (Section IV.B).
c. Risks (lethality) predicted to occur in Table
2 for pentachlorobiphenyl and hexachlorobiphenyl isomers, and
decachlorobiphenyl may not occur at some of the lower streamwater
concentrations (1.8 to 0.56 ug/1). The acute toxicity of
pentachlorobiphenyl and hexachlorobiphenyl isomers was shown
(Section IV.A) to decrease relative to acute toxicity of the
tetrachlorobiphenyls. It is possible that the chronic toxicity
for these isomers could also be less, however, under chronic
exposures much more time is available to take up and accumulate
an effective dose than under acute exposure conditions.
12
-------
2. Effect of Discounting Factors
a. Monochlorobiphenyls
Using a discounting factor of 50 for discharge of
monochlorobiphnyl isomers to receiving streams, will result in
(1) acute lethality to aquatic invertebrates at 30% of plants
during low streamflow, and (2) sublethal effects at 40% of plants
during low streamflow, and 10% of plants during average flow.
Acute lethality will be much greater than 50% at the 10
percentile (low flow), greater than 50% at the 20 percentile (low
flow), and less than 50% at the 30 percentile (low flow, Table
2). The net effect of the discounting factor will be to (1)
introduce acute risk to 30% of plant sites, (2) increase risk of
sublethal effects from 10% of plants to more than 40% of plant
sites at low flow, and (3) introduce potential risk from
sublethal effects to 10% of plant sites during average flow.
b. Dichlorobiphenyls
Use of a discounting factor of 5 for the discharge
of dichlorobiphenyl isomers to receiving streams will result in
acute lethality to aquatic invertebrates at 20% of plant sites
during low streamflow: lethality of greater than 50% will occur
at 10% of sites and lethality of 50% or less will occur at
another 10% of sites. Sublethal effects will occur at 30% of
sites during low streamflow. The result of using a discounting
factor of 5 will be to (1) introduce acute lethality and (2)
increase the occurrence of sublethal effects from 20% to 30% of
plant sites.
D. Risk Assessment for Algae
A risk assessment for algae cannot be done at this
time. The only toxicity information available is for marine
phytoplankton communities (Section V.A) and the hypothetical exposure
assessment is for freshwater riverine ecosysems (Versar 1983).
13
-------
E. Risk Assessment for Protozoa
1. Discharges of 100 ug/L
The risk of various PCB isomers to protozoa appears
to be low. The NOECs of 13 PCB isomers ranging from
monochlorobiphenyls through hexachlorobiphenyls were about 100
ug/L or greater. In the exposure assessment, the highest surface
water concentration estimated was 66 ug/L in streams with the
smallest dilution factor at low streamflow. This worst case
exposure condition is lower than the NOECs estimated for protozoa
(Section V.B)
2. Effect of Discounting Factors
a. Monochlorobiphenyls
Use of a discounting factor of 50 will introduce
risk to protozoa at 20% of plant sites during low streamflow
conditions. At 20% of sites, protozan growth could be reduced
50% in 43h (Table 6); at 10% of sites, growth could be reduced
more than 50%.
b. Dichlorobiphenyls
Use of a discounting factor of 5 for the
dichlorobiphenyl isomers could introduce risk to protozoa at 10%
of plant sites during low streamflow. At these sites growth
could be reduced about 10% to 50% in 43h (Table 6).
III. Toxicity of Various PCB Isomers To Fish
A. Monochlorobiphenyl Isomers
1. The subchronic no-observed-effect concentration
(NOEC) for monochlorobiphenyl isomers using data for the most
toxic isomer to the most sensitive fish species tested is
-------
estimated to be 50. - 80. ug/L to juvenile fish after about a 30-
day exposure.
a. The most toxic isomer of monochlorobiphenyl is
2-chlorobiphenyl. Dill et al. (1982) tested all three
monochlorobiphenyl isomers to three species of freshwater fish
(Table 3). The 2-chlorobiphenyl was the most toxic to all
species.
b. The most sensitive species to 2-chlorobiphenyl
was rainbow trout (Table 3).
c. The only NOEC for the monochlorobiphenyl
isomers is derived from a 32-d toxicity test for fathead minnows
to 2-chlorobiphenyl (Dill et al. 1982). The 96-h LC50 for
rainbow trout (540 ug/L) is about seven times lower than the 96-h
LC50 for fathead minnows (4000 ug/L, Table 3). Therefore, the
NOEC for fathead minnows was divided by seven to estimate a NOEC
for rainbow trout (i.e., 380. - 550. ug/L divided by 7 equals
50. - 80. ug/L).
2. The NOEC for the early life stages (embryo-sac fry)
of fish is estimated to be 2. - 3. ug/L.
a. Broyles and Noveck (1979_a_) reported that
several investigators (Schimmel et al. 1974, Nebeker et al. 1974,
and a personal communication from Mac M.J., W.H. Berlin, and D.V.
Rottiers, Great lakes Fisheries Laboratory, Fish and wildlife
Service, U.S. Department of Interior, Ann Arbor, MI. 48105)
"observed that fish of early developmental stages are more
sensitive to PCBs". Mac et al. indicated that the "highest
number of mortalities occurred before and up to yolk absorption;
fewer mortalities were observed thereafter" as reported by
Broyles and Noveck (1979^_).
b. Sac fry fish appear to be about 25 times more
sensitive to PCBs than juveniles or adults. Schimmel et al.
15
-------
(1974) found that Aroclor 1254 was 32 times more toxic to sac fry
than to juveniles of the sheepshead minnow (Cyprinodon
variegatus). Nebeker et al. (1974) found that Aroclor 1242 was
20 times more toxic to newly hatched fry of fathead minnows than
to 3-mo old fish.
c. Toxicity data for Aroclors were used to
supplement data for individual PCB isomers because it was known
that PCBs are more toxic to fry fish than juvenile fish, and the
only studies which measured the relative sensitivities of these
two life stages used Aroclors. These data for Aroclors are the
best estimates available for the relative sensitivities of fry
and juvenile fish and will be used for all PCB isomers classes
until data for isomers becomes available.
d. The NOEC for embryo-sac fry fish was obtained
by dividing the NOEC for the most sensitive fish (see Section
III. A.I above) by 25.
3. The acute (96-h) LC50 to juvenile rainbow trout for
the monochlorobiphenyl isomers is about 780 ug/L (Figure 2).
B. Dichlorobiphenyl Isomers
1. The 30-day NOEC for the dichlorobiphenyl isomers to
juvenile fish is estimated to be 12. ug/L.
a. No data were available for the
dichlorobiphenyl isomers (Table 3). Therefore, the NOEC was
estimated through statistical regression analysis of the
relationship between PCB chlorine number and available NOECs for
rainbow trout (Figure 3).
b. Figure 3 shows that as PCB chlorine number
increases, the bioconcentration potential of a PCB, as indicated
by it's octanol-water partition coefficient (Kow), also
increases. Figure 3 also shows that the toxicity of PCB isomers
-------
increases, as indicated by decreasing NOECs, with increasing
chlorine number.
c. It is assumed that chronic toxicity is
directly related to a chemical's Row, if the log Row is less than
about 6 or 7 and if the chemical is a non-reactive non-
electrolyte organic chemical. Heonens (1982) has shown that 16-d
ECSOs for Daphnia magna reproduction (a chronic toxicity
endpoint) are linerally related to log Kow for a variety of
organic chemicals. Konemann (1981) has shown a strong
relationship between 14-d LCSOs for guppies and log Kow (up to a
log Kow of 6) for a group of non-reactive, non-electrolyte
organic chemicals.
2. The NOEC for the dichlorobiphenyl isomers to the
early life stages (embryo-sac fry) of fish is estimated to be
about 0.5 ug/L.
a. The rationale is the same as presented above
in Section III.A.I.a through d.
3. The acute (96-h) LC50 to juvenile rainbow trout for
the dichlorobiphenyl isomers is estimated to be about 420 ug/L
(Figure 2).
C. Trichlorobiphenyl Isomers
1. The 30-day NOEC for the trichlorobiphenyl isomers
to juvenile fish is estimated to be about 2.1 ug/L.
a. A chronic NOEC was not available for the
trichlorobiphenyl isomers (Table 3). Therefore, a NOEC was
estimated through regression analysis of the relationship between
toxicity and PCB chlorine number (Figure 3).
b. The rationale for using Figure 3 is the same
as presented in Sections III.B.l.b and c.
11
-------
2. The NOEC for the trichlorobiphenyl isomers to the
early life stages (embryo-sac fry) of fish is estimated to be
about 0.1 ug/L.
a. The rationale is the same as presented above
in Sections III.A.I.a through d.
3. The acute (96-h) LC50 to juvenile rainbow trout for
the trichlorobiphenyl isomers is estimated to be about 220 ug/L
(Figure 2).
D. Tetrachlorobiphenyl Isomers
1. The 30-day NOEC for the non _o,_p_'-chlorine (CD
substituted (Figure 1) tetrachlorobiphenyl (TCB) isomers (5
isomers out of 42 possible isomers) to juvenile fish is less than
0.1 ug/L.
a. Stalling et al. (1979) have demonstrated that
the NOEC for rainbow trout exposed to 3,3',4,4'-
tetrachlorobiphenyl (TCB) for 50 d was less than 0.1 ug/L which
was the lowest exposure concentration tested (Table 3).
b. Stalling et al. (1979) reported that studies
by Goldstein et al. (1977) and Poland and Glover (1977) concluded
that PCB isomers lacking _p_,^'-chlorine (CD substitution and
having four or more Cl atoms, may account for a significant
amount of the toxicity of PCB mixtures.
c. Stalling et al. (1979) determined the
toxicities of four groups of PCBs: (1) 3,3',4,4'-
tetrachlorobiphenyl (a Oc^jo'-Cl substituted PCB), (2) a mixture
of 1 _0'_o'~cl PCBs, (3) a mixture of 1 and 2 _o,^'-Cl PCBs, and (4)
a mixture of 2 - 4 o_,o_'-CI PCBs. The 3 , 3 ' , 4 ,4 '-TCB was more
toxic to rainbow trout than any of the 1 thru 4 _p_,_o_'-Cl mixtures.
-------
d. Bruggeman et al. (1981), Shaw and Connell
(1980_b), Goldstein et al. (1977), Poland and Glover (1977), and
Stalling et al. (1979) have all discussed the importance of the
ortho-ortho chlorine (jo^^'-Cl) substitution pattern within a
class of PCB isomers and its effect upon bioconcentration
potential or toxicity. In general, it is suggested that the more
chlorines substituted in the ortho-ortho positions on the
biphenyl, the lower the bioconcentration potential and the
toxicologic activity. Ortho substitution of Cl forces the
biphenyl out of a common plane. In summary, the toxicity of PCBs
appears to increase with the number of chlorines (Figure 3), and,
within a class of PCB isomers, may decrease with greater ortho-
ortho Cl substitution.
2. The 42-day NOEC for the 1 thru 4 jo,o_'-Cl
substituted TCB isomers (37 isomers out of 42 possible isomers)
to juvenile fish is about 1.5 ug/L.
a. Branson et al. (1975) have demonstrated a NOEC
for rainbow trout exposed to 2 , 2 ' ,4 , 4'-TCB (a 2_o,_o'-Cl
substituted TCB; one Cl on each side of the biphenyl bond; Figure
1) for 42 d was 1.5 ug/L (Table 3). Branson et al. also showed a
no-observed-lethal concentration (NOLC) of greater than 14
ug/L. These data are supported by the data reported by Dill et
al. (1982) who exposed fathead minnows to 2,2',4,4'-TCB for 30
d. Dill et al. reported an LC50 of 29 ug/L and a NOEC of less
than 14 ug/L (Table 3).
b. These data suggest that a 0 _o,_o'-Cl
substituted TCB is more than 15 times more toxic to juvenile and
adult fish than a 2 _o,_o_'-Cl substituted TCB.
3. The NOEC for the 0 _o,_o'-Cl substituted TCB isomers
to the early life stages (embryo-sac fry) of fish is estimated to
be less than 0.004 ug/L.
-------
a. The rationale is the same as presented above
in Sections III.A.I.a through d.
4. The NOEC for the 1 thru 4 _o,^'-Cl substituted TCB
isomers to embryo-sac fry fish is estimated to be about 0.06
ug/L.
a. The rationale is the same as presented in
Sections III.A.I.a through d.
5. The acute (96-h) LC50 to juvenile rainbow trout for
the l-4-_p_,_p_'-Cl substituted tetrachlorobiphenyl isomers is
estimated to be about 120 ug/L (Figure 2, Table 3).
a. The acute (96-h) LC50 is for 2,2'2,4'-
tetrachlorobiphenyl (Table 3).
6. The acute (96-h) LC50 to juvenile rainbow trout for
the non _0_f.O_'-Cl substituted TCB isomers could be as low as 10
ug/1.
a. It is probable that the acute toxicity of non
_p_,_0_'-Cl substituted TCB isomers to juvenile fish is greater than
the acute toxicity of the three 4 _p_,_0_'-Cl substitute isomers
because the chronic toxicity of the non _p_,_p_'-Cl substituted
isomers was greater than the 1-4 _0_,_p_'-Cl substituted isomers by
about 15 times.
b. The acute (96-h) LC50 for the 1-4 _0,_p_'-Cl
substituted isomers was devided by 15 to estimate an acute (96-h)
LC50 for the non _p_,_0_'-Cl substituted isomers.
E. Pentachlorobiphenyl Isomers
1. The 30-day NOEC for the pentachlorobiphenyl isomers
to juvenile fish is estimated to be 0.07 ug/L.
20
-------
a. No data were available for the
pentachlorabiphenyl isomers (Table 3). Therefore, the NOEC was
estimated through regression analysis of the relationship between
toxicity and PCS chlorine number (Figure 3).
b. The rationale for using Figure 3 is the same
as presented above in Sections III.B.l.b and c.
2. The NOEC for the pentachlorobiphenyl isomers to the
early life stages (embryo-sac fry) of fish is estimated to be
about 0.003 ug/L.
a. The rationale is the same as presented in
Sections III.A.I.a through d.
3. The acute (96-h) LC50 to juvenile rainbow trout for
the pentachlorobiphenyl isomers is estimated to be similar to or
less than the LC50 for tetrachlorobiphenyl isomers (about 120
ug/L).
a. The acute toxicity (96-h ECSOs) of PCB isomers
to aquatic invertebrates decreases with the higher chlorinated
isomers (pentachlorobiphenyls and hexachlorobiphenyls). This
decrease in toxicity is probably associated with the decreasing
water solubility of these higher chlorinated PCBs. A more
detailed rationale is in Section IV.A.
b. Since no toxicological information is available
which indicates how much less acutely toxic the pentachloro-
biphenyl isomers are to fish than the tetrachlorobiphenyls, acute
toxicity was assumed to be similar.
F. Hexachlorobiphenyl Isomers
1. The 30-day NOEC for the hexachlorobiphenyl isomers
to juvenile fish is estimated to be 0.01 ug/L.
21
-------
a. No measured NOECs are available for the
hexachlorobiphenyl isomers (Table 3). Therefore, the NOEC was
estimated through regression analysis of the relationship between
toxicity and PCB chlorine number in Figure 3.
b. The rationale for using Figure 3 is the same
as presented above in Sections III.B.l.b and c.
2. The NOEC for hexachlorobiphenyl isomers to the
early life stages (embryo-sac fry) of fish is estimated to be
less than 0.001 ug/L.
a. The rationale is the same as presented above
in Sections III.A.I.a through c.
b. In addition, Shimmel et al. (1974) exposed
fry, juvenile, and adult sheepshead minnows to 0.1 ug/L Aroclor
1254 and demonstrated that fry were about 30 times more sensitive
to PCBs than older fish. These data suggest a NOEC of 0.003 ug/L
(i.e., 0.1 ug/L divided by 30) and support the estimated NOEC for
sac fry fish of about 0.001 ug/L hexachlorobiphenyl.
c. The 2,2',4,4',5,5'-hexachlorobiphenyl (HCB)
isomer (a 2 o_,o_' -Cl substituted isomer, one Cl on each side of
the biphenyl bond) has been shown to be very toxic to sac fry of
lake trout and Chinook salmon (Table 3). Broyles and Noveck
(1979b_) exposed lake trout sac fry to 8 ug/L 2, 2 ' , 4 , 4 ' , 5 , 5'-HCB
for 15 d; all fish died within 79 d. All Chinook salmon sac fry
died within 31 d.
3. The acute (96-h) LC50 to juvenile rainbow trout for
the hexachlorobiphenyl isomers is estimated to be similar to or
less than the 96-h LC50 for tetrachlorobiphenyl isomers (about
120 ug/L).
22
-------
a. The rationale is the same as for pentachloro-
biphenyls (Section III.E.3).
G. Decachlorobiphenyl
1. The 30-day NOEC for decachlorobiphenyl to juvenile
fish may be similar to or less than the hexachlorobiphenyl
isomers. Since the chronic toxicity of decachlorobiphenyl to
juvenile fish is not known, it will be assumed that toxicity will
be similar to the hexachlorobiphenyl isomers: about 0.01 ug/L.
a. No data were available for decachlorobiphenyl
(Table 3) and a NOEC was estimated through graphic interpolation
of Figure 3. Estimating a NOEC for decachlorobiphenyl from
available data or other PCB isomers will have a high degree of
uncertainty due to decachlorobiphenyl's greater water
insolubility relative to PCB isomers with six or less chlorines.
b. Konemann (1981) has shown a deviation from a
linear quantative structure-activity relationship between log Kow
and subchronic toxicity (14-d LC50) to fish with chemicals whose
log Row's are greater than six. For such chemicals effective
concentrations should be larger than expected from the struture-
activity relationship, and, thus, less toxic. However, the
National Research Council (NCR 1979) reported that the degree of
bioaccummulation of decacholorobiphenyl in aquatic organisms is
greater than for 2,2',4,4',5,5'-hexachlorobiphenyl. Factors of
97,000 in fish and 930,000 in snails were measured in the
terrestrial aquatic model ecosystem of Dr. R. Metcalf. Factors
for hexachlorobiphenyl were 42,000 in fish and 100,000 in
snails, If the bioaccummulation potential of decachlorobiphenyl
and the hexachlorobiphenyl isomers are similar, their toxicity to
fish could also be similar.
c. The simplist assumption is that the structure-
activity relationship becomes asymptotic. Sugiura et al. (1978)
23
-------
have suggested that bioconcentration factors become asymptotic
for chemicals whose log Row's are greater than six.
2. The NOEC for decachlorobiphenyl to embryo-sac fry
fish may be similar to or less than the estimated NOEC for
hexachlorobiphenyl: about 0.001 ug/L.
a. The rationale for using the value for
hexachlorobiphenyl is presented above in Sections III.G.I.a
through c.
3. The acute (96-h) LC50 to juvenile rainbow trout for
decachlorobiphenyl will be similar to or less than the LC50
estimated for tetrachlorobiphenyl isomers (about 120 ug/L).
a. The rationale for using the value for
tetrachlorobiphenyls is presented above in Section III.E.3.
IV. Toxicity of Various PCS Isomers to Aquatic
Invertebrates
Various PCB isomers are acutely toxic to aquatic
invertebrates and toxicity increases with an increase in the
number of chlorines on the biphenyl up through
tetrachlorobiphenyl. Acute toxicity ranges from 700 to 30
ug/L. Chronic toxicity of PCB isomers to aquatic invertebrates
is not available, however, aquatic invertebrates are expected to
have lower chronic NOECs than those measured for juvenile fish.
A. Acute Toxicity
1. Various PCB isomers are acutely toxic to aquatic
invertebrates. Toxicity increases with chlorination and reaches
a maximum with the tetrachlorobiphenyls at about 30 ug/L.
24
-------
a. Static acute (48-h or 96-h) toxicity values
(ECSOs) have been measured for one or more isomers of every PCB class
from monochlorobiphenyl through hexachlorobiphenyl (Table 4).
b. The EC50 values of the monochlorobiphenyl
isomers through the tetrachlorobiphenyl isomers are linerally
related (Figure 4) and this relationship can be defined through
statistical regression analysis: log EC50 (ug/L) = 3.04 - 0.411
chlorine no. (R2 = 0.92, N = 7). These data indicate that the
average EC50 value for the isomers of each class of PCB from one
to four chlorines are:
CLASS VALUE (ug/L)
monochlorobiphenyls 500.
dichlorobiphenyls 110.
trichlorobiphenyls 70.
tetrachlorobiphenyls 30.
2. Acute toxicity of higher chlorinated PCBs (five or
more chlorines) is reduced due to their water insolubility.
a. Isomers of higher chlorinated PCBs, i.e.,
pentachlorobiphenyls, hexachlorobiphenyls, and decachlorobiphenyl
are less acutely toxic than tetrachlorobiphenyl isomers (Table 4,
Figure 4). These higher chlorinated isomers are becoming
relatively more water insoluble which results in reduced uptake
rates into aquatic organisms. Thus, acute exposure to low water
concentrations does not permit enough PCB to be taken up to kill
50% of a test population. Konemann (1981) and Sugiura et al.
(1978) have observed reduced acute toxicity and bioconcentration,
respectively, with decreasing water solubility of organic
chemicals whose log Row is greater than six.
B. Chronic Toxicity
1. Chronic toxicity NOECs are not available for
25
-------
individual PCB isomers, however, NOECs for aquatic invertebrates
are expected to be equal to or less than the subchronic NOECs
measured and estimated for juvenile fish (Table 5).
a. Aquatic invertebrates have been shown to be
more sensitive (i.e., lower EC50 values) to various PCB isomers
than fish during acute exposures (Mayer et al. 1977; and Tables 3
and 4) or just as sensitive as the most sensitive fish species
tested (Dill et al. 1982, and Tables 3 and 4).
b. It is assumed that trends between aquatic
invertebrates and fish with regard to acute toxicity of PCB
isomers will also be observed with chronic toxicity information
when it becomes available.
c. A NOEC of Aroclor 1254 (i.e., "a safe level of
A-1254") for an aquatic invertebrate was suggested as being below
1 ug/L by Nebeker and Puglisi (1974). They calculated a 3-wk
LC50 of 0.45 uy/L of Aroclor 1254 with respect to reproductive
impairment. This NOEC suggested for Aroclor 1254 is consistent
with NOECs for PCB isomers to juvenile fish (Table 5).
V. Toxicity of Various PCB Isomers to Algae and Protozoa
A. Algae
The PCB isomers may be just as toxic to algae as they
are to juvenile fish.
1. Only 2,4'-dichlorobiphenyl (DCB) has been tested.
Moore and Harriss (1972) predicted a NOEC of less than 7 ug/L
after a 24h in situ exposure of 2,4'-DCB to a natural marine
phytoplankton community (Table 6). This NOEC is similar to the
30-day (subchronic) NOEC estimated for the DCB isomers to
juvenile fish (12 ug/L, Section III.B.I, Figure 3).
26
-------
B. Protozoa
The NOECs of PCS isomers (monochlorobiphenyl through
hexachlorobiphenyl) to protozoa appear to be about 100 ug/L or
greater.
1. Dive et al. (1976) measured the effect of 13 PCB
isomers (Table 6) on the growth of the ciliated protozoa
(Colpidium campylum). In general, Dive et al. concluded that
toxicity decreased as the number of chlorines increased.
2. Dive et al. also argued that their results with
isomers were consistent with results from Aroclor mixtures using
several protozoan species. With one exception, all NOECs for
Aroclor mixtures were 100 ug/L or greater.
VI. Toxicity of PCB Isomers to Wild Mammals and Birds
Toxicity studies of PCB isomers to wild mammals and birds are
not available.
VII. Environmental Concerns of Polychlorinated Biphnyl (PCB)
Commercial Mixtures
Ambient .concentrations and food chain transport of PCBs may
impair the reproductive potential of commercial fisheries and
wild mammals, e.g., mink. PCB residues are also strongly
correlated with reductions in natural populations of marine
mammals and may be correlated with declines in river otter
populations. High PCB residues have been found in various birds,
especially gulls and carnivorous birds, but no resulting effects
have been firmly demonstrated.
A. Commercial Fisheries
PCBs may, even at low concentrations, contribute to a
27
-------
reduction in populations of sport and economically important
fish.
1. PCBs may affect the reproductive success in fish.
a. High body burdens (120 ug/g wet weight in the
ovaries) in wild fish (Baltic flounder) been correlated with
failure of eggs to hatch. (Von Westernhagen et al. 1981).
b. Experimental data have demonstrated that PCBs
can reduce spawning, hatching, and survival of many species of
fish. Of the species tested, one (brook trout) is important to
recreational fisheries. Bengtsson 1980; Mauck et al. 1978; DeFoe
et al. 1978).
c. Seelye and Mac (1981) exposed fry hatched from
eggs from Lake Michigan lake trout and fry hatched from hatchery
lake trout to 50 ng/L Aroclor 1254 for 50 days. The observed
mortality was site-specific. Only the Lake Michigan fish, both
the exposed fry and the unexposed control fry, showed size-
specific mortality, i.e., the smaller fry died first. The Lake
Michigan fish had large residues of PCBs prior to testing; the
hatchery reared lake trout had relatively small PCB residues.
Although several factors could cause size-specific mortality,
Seelye and Mac (1981) suggested that the cause was the large PCB
residues.
d. Early life stages (embryos and sac fry) of
fish appear to be about 25 times more sensitive to PCBs than
juvenile and adult fish (Schimmel et al. 1974, Nebeker et al.
1974) .
e. Predicted no-observed-effect concentrations
(NOEC) for the early life stages of fish can be found in Table 5
for each class of PCB isomers and these NOECs are similar to
toxicity data for PCB commercial mixtures.
28
-------
f. Many fish species spawn in shallow near-shore
areas, and eggs and sac fry rest on the top layer of sediment
where PCBs tend to accumulate in the aquatic environment. The
early life stages of fish have the potential to receive high
exposure of PCBs.
2. PCBs may have direct adverse effects on juvenile
and adult fish.
a. Experimental laboratory data have demonstrated
that both growth and survival of many fish species are reduced at
very low exposure concentrations (0.4 and 4.0 ug/L,
respectively).
b. Mauck et al. (1978) exposed brook trout fry to
Aroclor 1254 for 118 days post-hatch. Growth was reduced at 1.5
ug/L while survival was reduced at 3.1 ug/L.
c. DeFoe et al. (1978) performed a 240-day life
cycle test with fathead minnows and Aroclors 1248 and 1260. The
lowest effect concentrations were 0.4 ug/L Aroclor 1248 for
reduced growth and 4.0 ug/L Aroclor 1260 for reduced survival.
d. Toxicity of PCB Aroclors and PCB isomers to
fish appears to be related to the number of chlorines attached to
the biphenyl ring (Figure 1). The greater the number of
chlorines (up through six chlorines), the greater the toxicity
(Table 3, Figure 3). This toxicity appears to increase an order-
of-magnitude with each added chlorine.
e. Subchronic NOECs for juvenile fish can be
found in Table 5 for each class of PCB isomers.
f. PCB isomers and mixtures that have no o_,o_' -Cl
substitution appear to be more toxic than PCB isomers with 1 to 4
chlorines substituted in the _o_,_o' positions on the biphenyl
29
-------
(Figure 1). See information for tetrachlorobiphenyl in Section
III and Table 3).
g. Impaired bone development and abnormalities in
testes have been observed in two species of fish (brook trout and
Atlantic cod) fed or exposed to low levels of PCBs (0.4 ug/L).
(Mauck et al. 1978; Sangalang et al. 1981).
3. PCBs can affect the survival of aquatic
invertebrates and aquatic plants which are food sources for the
fish.
a. PCBs can have lethal and sublethal effects on
environmentally important freshwater invertebrates.
(1) Experimental laboratory data have shown
that PCBs are toxic to many aquatic invertebrates (Daphnia magna;
juvenile scuds, Gammarus pseudolimnaeus; and midges, Tanytarsus
[Paratanytarsus] disimilis) in the low ug/L range. (Nebeker and
Puglisi 1974).
(2) Experimental data have demonstrated that
very low concentrations of PCBs can result in reproductive
impairment of aquatic invertebrates (Daphnia magna reproduction
was impaired by 16% with 0.48 ug Arclor-1254/L and Gammarus
Pseudolimnaeus reproduction was impaired by 50% with 5.1 ug
Aroclor- 1248/L). (Nebeker and Puglisi 1974).
b. PCBs affect productivity of phytoplankton and
the composition of phytoplankton communities.
(1) PCBs (1 ug/L) decrease the photosynthetic
rate of different species of algae. (O'Connors and Mahanty 1979,
Kricher and Bayer 1977).
30
-------
(2) PCBs are differentially toxic to
different species of algae at or below 1 ug/L. (O'Connors et al.
1978, and Glooschenko and Glooschenko 1975).
B. Contamination of Food Resources
1. PCBs can be concentrated and transferred in fresh-
water and marine phytoplankton, invertebrates, fish and mammals;
and can result in indirect human exposure by consumption of
economically important food resources, the closure of fisheries,
and economic losses as a result of this contamination.
a. Numerous sources have demonstrated and
reviewed the propensity of PCBs to bioconcentrate in aquatic
organisms (32,OOOX-270,OOOX) and to be transferred upward in the
food web. (Keil et al. 1971; Biggs et al. 1980; Anonymous 1980;
Shaw and Connell 1980j|_,_b; Bleavins et al. 1980; Thomann 1978;
Weininger 1978; Peterson and Guiney 1979; US EPA 1977, 1980).
b. Residue data collected in the environment
demonstrate that PCBs are ubiquitous in aquatic organisms and can
be bioaccumulated as much as 15000X over short periods of time
(14 days) from contaminated sites based on water
concentrations. (Skea et al. 1979, Anonymous 1980, Swain 1980,
Davis et al. 1981, Risebrough et al. 1968).
c. Polychlorinated biphenyls can also be
bioconcentrated in fish and aquatic invertebrates through
contaminated diets and diets alone may contribute a significant
proportion to the total body residue of PCBs. (Pizza and
O'Connor 1983, Spigarelli et al. 1983, Rubinstein et al. 1983).
d. Residue data collected from commercial and
sport fisheries demonstrate that significant quantities of PCBs
may be transferred to humans through consumption of fish. (Swain
1982, Schmitt et al. 1981, Zimmerman 1982).
31
-------
C. Mammals and Birds
1. PCBs may impair reproductive success in some wild
mammals (e.g., mink) and birds.
a. Clinical signs (e.g., reproductive failure,
death of female breeder mink, impaired growth of kits, excessive
early mortality of kits, reduced birth weights of kits, reduced
litter sizes, emaciation, anoroxia, and blood stools), pathology
(e.g., liver enlargement, hemorrhagic ulcers, degeneration of the
liver and kidneys, and fatty infiltration of the liver), and
mortality patterns are very similar for mink fed (fish) diets
containing PCBs and for mink fed PCBs. (Aulerich et al. 1977).
b. Dietary exposure to mink to PCBs (5 mg/kg)
causes reduction in number of kits and reduced survival of
kits. (Bleavins et al. 1980, Aulerich et al. 1977, Jensen et al.
1977) .
c. PCBs accelerate destruction and alter
biosynthesis of normal body steroids (estradiol, testosterone,
rosrene-3,17-dione, pregnenolone, progesterone, and
androstenedione) in birds and wild mammals. Hormonal alteration
can affect mammals that exhibit delayed implantation and mating
behavior in birds. (Risebrough et al. 1968, Lincer and Peakall
1970, Nowicki and Norman 1972, Freeman and Sangalang 1977,
Reijnders 1980).
d. PCBs are correlated with reductions in natural
populations of marine mammals, such as, harbor seals, ringed
seals, and California sea lions. (Helle et al. 1976; Reijnders
1980).
e. PCS residues higher than those found in
reproductively impaired mink have been found in river otters in
the lower Columbia River Valley in Oregon. River otter harvests
32
-------
by trapping have declined in this area, but have risen in other
parts of Oregon where PCB residues are negligible or undetectable
(Henny et al. 1981 ) .
f. Experimental laboratory studies have
demonstrated that PCBs in the diet (10 mg/kg) of birds (i.e.,
ring doves and ring-neck pheasants) caused early embryonic
mortality and reduced egg production. (Peakall et al. 1972;
Dahlgren and Linder 1971).
g. High PCB residues have been found in a number
of avian species, especially in the Great Lakes area. Some birds
(e.g., gulls) have shown population declines, but no cause and
effect relationship has been determined. (Heinz et al.
Unpublished Manuscript, Gilbertson and Hale 1974).
33
-------
: -o £ o c
I 0 a 1 0 •*
N -H JJ B
• C-H vi wi e
4) U «J «O O
i 0 o jj a c -H
' 4J ^J ^^ Q 4*
i £ o 3 0 o
) en a «a as
< i O i 4)
£ -O U M
: o c "o ja
3 • «J 4) E n
t B«O JJ 4)
t h c o o o
3 4) C r-l 4) -
•( S -H VI •
J Q -a C VI <•* B
} a 3 o ••*
: a 0 -f-i jj -vi
3 O B £
j a« B o o a jj
«i i 1 1 ^j ^j
•M *• *^ ^^
4 «-t VI 3
i IB o >co -a
J 3 V* JJ 0 VI 0
1 -O 4) -* JJ O
t-H * ^ O t>
> « a) a e
(•H — £ CU0 «
j >O n JJ x o>
0 COB 0000
U| ^ O» 1-1 JJ -H
.1 -H e a -H
3 u i o e
n o u •* « 0
vi « >j jj vi >
H a « -* 3 •
i 0 h ^ i-t t-iJJ
j r-t 0) • 3 U
Hd>ao.>iO0
u B— e g >* J» vi*
V »H O CU W M
: «« ,H-H « a 0
U B 1 JJ « « JJ
O > S 0 O ^
oL« ^ vi 0 0 a
>i 0 3 O J3 VI £
c a ^ a jj «w jj
41 JJ O JJ 00
0 > «J Q.C O "•
U H 4) JJ ^ JJ
US £ O 03
^ O S ® fc4 O J3 O
0 -H 0 JJ JJ
W0WViCUO0iJ
3 a vi 0^0
U C O B i C
-tQ^OJJUUW
ce o>£ vi — -H vi cj>
5
^
I ~*°
|
««
•
h«
i "
c S
,. z
• —
5 8
li
H
?f
£ e
? 0
i i
1 s
;i
K
•
i
-OT
-9
-S
.o^o o -»
'o >-T -t
-e
-r
-T
=t 1
1 i
i i
3 8
S f
««
j
u
§
b_
-T '
3^
i -
s i
£4
1
J
1!
S 1
s
^ >
•• 0
« C
? 1
3 *
8
«
L.
^
«•
§]
^
I;
!l
|v ^ | 1 1 |-*J-^i— J-^]-«sl-^J
^4\^-^j^J-J^J vo^^o\o
-------
10
u]
u)
ui
Uj
w
§
I
i
S
§
s
1-
Ml
PI
fll
i4i
-g
1 Fi
i si
1 !>
IB
f -5
i I|
? I *
35
-------
•Q
o
09
09
rH 4> I
§«
00 |J
•O B
I
o
a
a
u
C-H
0 -W
> 4) • *> £
•H 14 00 B O1
•o 4) ««-i a
C **> 3
!•* B a o
i —. H O -W
I fc4 f*> J9 Ci4
I Q 00 0) 09
VI Wl JJ « JJ
rH U U
14) 0) VI 4)
^1 U > O VI
i a B c vi
i B n-H JJ 4)
•H wi e
•H 4) O 41 <-l
> B >*H O B
> ~-JJ M £
B B at JJ
pH 3 Qi 4)
oj i a* i-i
. « a « » A
i > 4) 3
U 4} V| 4J 00
3 -H O B
I U £ o 00
B «-H 4)
I 4) EH C-O -U
4
oc
•
i o B *> -H
i tn vi ^ o. c
10) 3-H -H
I U B D.U
i c o o n
I 4) O O. fl)
I O -H .Q tJ
O -U O 3 W
' -a B -u a
t°
••D
B 4) «
J 4) JJ 00 » •
} 45 -H 4V» JS *-
H jj n o JJ 4J
V «— o
8 VI 0) , a 4) w o
6 C SJJ C-H
« -ff n t» JJ
C.C M
a 4) « 14 o.
U -O 4) O O
J 4) JJ VI -H B
B 9 £ JJ
4 u OITI a N
a x~* « s ja
I) 4) H N CT
J -H i) "B •
4 C 1-t 0) 09
1 O • « O JJ 0)
• M B JJ JJ O *J
: jj t< o o> «
H B 4) B >*IH M
14 g Vi -M VI U3
J JJ O «-« « 0)
Ft C 09 J=^ JJ
W 4)-H O g $ Vj
B U J= A «
3 C ffl -O JJ >
cr o u e 4> o e
B O Q. B^l JJ-H
-,o*o 0
-£
-r
-I
-J
-J
vo
vn
VO
S 4
a s
§
* 8 8 9
8 9
36
-------
(n
to
-J
§
*
g
§
s
:«>
i-!i
Hi
\*~
- s "
*f*
if!
**S
Mi
«.r-.
III!
s
I
I §t
i 11
* 1|
I If
28
1
|
i
L.
11
li 1 i
S3 S is =
f'* - ^
37
-------
0!
tt
0
a
•c
0
*j
1C
c
**4
u
C
£
U
£
(0
10
« 3
u
e
a
cr.
cc
9
(N
CC
(N (M
z
cc
-i a
m Q
3 0
-H ij
(E —
*2
o a
c n
« o
10 a
b. —
ȣ
o a
c oi
a
10 a
e 0
25
U
*s
§S
*•"«.
**
II
-4 «
0)
3
*
a
» a
o —
c u
C 10
•c o
c -c
0g
a u
a> a
0 >
£ U
to ^
n
a
0
,2
00^
c n
x-H
a
•c £
10 a
o
r
c .
10
K •
01
IB a
I
cs
I
(N
-------
c
o
u
t
M
0)
S
(Q
0
2
£
o
UJ
a
IX
e 2
"Je 3
u
*u
14j
cc
*
•c
0
JJ
0
x
a
01
u
ID
Q.
to
U
O
0
to
M
M P»l
e ec
•-I **• r*
oe
• • a-
* c
c
jj jj C
o e 2
^ ~ at
-* ~* c
» — o
c c *
ee oeeeecee
c cc cinocoee-ircc
ffVC rtfO^l^W^^i-*
A
e e o
in IT in
S U i
Z -H Z
ca a u
K > S
ft, m o,
H^ W
ffl (D A
O M O
e m a;
S . 3 ,
x *j = 5
£ jc o a o Z
S S M "O M »
* » Q o g z
«, «: O r* *,
Z Z 2 OS
te » *>
en x co
n
3 *-
4J IB
Ol IB
•« Q -^ 3
U C U U
jj c «M ^ JJ
§^3 Z ®
u c u
ki -< TO
E- IB «J -C «
CT » O *•»
J1 £ e rt
Q » »• —
law > o
C pH O Q 0U Q}
— i m a >• Q. o
-------
g
o
a
a
a
?
ft
-H O-
IB O-
JJ —
cr
a
c
0
tt Ifl
c r-
a 9
e c ^
u
0)
11 c
c c -S c
.. 0 C JJ O
•U jj jj ce E JJ
c o o *• r» 3 co
tf 3 3 >."C Bl • JJ
^ "C •c jj c -c c
JJ 10 3 t-i
R JJ
g*
hi JJ
0 n
D> 0
3
3
td
ft.
M
ffl
§
O
J
a
§
H
Cd
H
in c
c 5
u
r- 3
•c
0)
_. JJ U
CJ -C 3 0) <*•
61 11 O > «
C ffl 1J -<
-------
JJ
c
o
o
*
f*>
0)
•-4
£
<8
6-
9
g
a
u
8
a
ec
a u
f«* m
•0 3
±j
U
a>
tu
IU
41
•C
0
JJ
•-
Q) >« JJ JJ Q} IQ CO O*
UJ 4J IQ C UJ 0) 4)
MJ —4 N 9) vw > UJ OiJ:
(D C |a JJ —4 ki
^ £ (Q C ^ "** cfi O1 C
eucjj**tea)a>a:.- £ - Z
a- a- z w
i • . a.
z z «
^M »fc * *
d. CR c/: en
tn n 0
5 Ji «
1 ^4 Us *3
0 la U U
3 la . ve
^* « » ^(* *
. . . » .
CN CN CN m CN
« » te k «fc
CN CN CN CN (N
d)
a
a
•H
(Q
>
«
O
U
a
a
a
V
0
z
-------
•
JJ
e
0
o
•
po
o
£
IB
H
0
g
0
I*
0
—
jj
O
0
iw
UJ
tfc
«
•c
0
£
JJ
0
ffl
O
•*H
U
0
a
K
£ £
» *
r r- -c i-
e 9 a 9
n ~ « —
a: j* « x
j. o 0 U
•« 0 i-i0
> > > >
00 00
u z )-> z
C (C
•c -c
a- -*
•• r- •• m
> ><
jj >. JJ >,
-I £ x £
10 «c IB *
£ C £ C
JJ C JJ C
. 5- 3-
J
>l
u
&* * ^T
H .2 « J
S ^^ >NV
2 i- o i- 01
g \w 3 «-i 3
d O OE O OC
=• IB " IB ~
9 « r cr. r
g a. M 0. w
>< -x J -x J
g r u u r K B
K -c -c aTr r
• in v «m \e
tn ^< x ec -H -j
a
£
— O
£ D!
« JJ
3 >
U »
>. «
U rH .* £
H 0 C u
> 00
0 •* CO
J^ B •* C
IB U £ O
J — O —
in in
^ ^
m m
v ««•
* ^
fM
DECACHLOROB
o da'ta are a
z
42
c
c
ffi JJ
C IE
C ^
-- JJ
JJ =
c o
w u
JJ C
c c
a u
O O
u 0
iw
•C
0 u
E 0
E
a £
o
".i
o u
u cc
£ C
JJ Z
* 5
1-1 0
II 3
—
H C
b. (C
0 0) 5
3 0
-* JJ £
B n
> «j o
0 u
jj iu
O -<
0 10 U
iu 3 0
Urf 0
0 C jj
0 0
0 1- S
JJ
5 II II
- K JJ
JJ 3
0! - X
0 •C
o -c
0 — C
JJ U IB
o
•c a -
0 c
a c c
3 C •*
-< JJ
ffl JJ B
C B U
O C jj
•*• •* C
JJ E 0
1C — U
U r* C
jj 0 O
C U
0 I
o f> »<
O J-H
U IB
C -O O
^ 0 r+
e -i i
C U -O
C 0 0
a >
« b
0 0
Z l-c ffl
3 .a
• BJ 0
O 0 I
•** a. o
JJ X C
B 0
JJ II
a R
u
» a. j
x o
MUZ
-------
03
0)
(0
u
dJ
±J
u
0)
c
M
o
(8
3
O1
O
JJ
01
CO
O
04
0)
c
0)
a
•*-4
03
T3
0)
4J
(0
c
••-4
^J
o
,-4
|
O
a.
U-l
0
01
0)
S
o
01
M
03
o
•ft
lJ
<0
UJ
o
JJ
0
X
8
^r
P_<
JD
«J
£-
4>
O
4)
U
4)
144
4)
as
**•*
4) J
3 \
«-*
JJ
U
4)
VM
Cd
*
TJ
o
JJ
4)
01
4J
•H
O
a
en
S
e
o
to
M
*~* ••••i •—• p* t**
(N tN (N f~ P»
OO 00 00 9\ 9\
~ ~ "* A . .
• • • —t ^ fH
^ ^ ^4 *^* • *0 ^0
<0 (0 <0
•O jJ JJ
JJ JJ JJ C 4) 4)
4J 4» 4) <0
i— ( Sj ti
rH r-l ^ 4) >1 >t
•^4 **^ **4 W ft flj
a a a > z £
vo
... a •
o o o • o o
O <** i
o - JJ
. O 4) -4
<«J QlrH
en m -C 'O *8 < in in *" «2 w-«4Qu4)ijm m
z cj o OS Oa)--4>ocj cj
S U U " § 2 2 J < E W U
a. £
i 1
0 § |0
XQOOOCO^ZVO vovo
CJ ^^ ^* ^* ri ^ ^^ ^^
O**fcuj2 » »
zzz^S^ ^^
s en en en &< en en
01 01
— » 33
U (0 (0
•-4 C C
c e e
••H --4 *H
ty *™4 ^H
U O O
<—» ^» ^^ i^ TJ T3
10 (0 <0 > 33
C C C 4> CO
O^ O^ D^ *0 01 01
(0 (0 *0 V Q< &
E e E u
JJ 01 01
CQlQoi'0'01'8 01 3 3
•Q -H »O -H «0 -H O i-l U
•HC-^C-^C ijco - a *- o^- to ^- en >—
i
i ~ •
ro ^f ' 'V
\ \ \ - *
CN f*1 ^» CN CS ^>
r*«
r»
2
.
r«4
(0
JJ
4)
ki
S,
(0
X
0
r^
2 o
>t in
Z 0
CU
CO
o
cc
o
3g vO
CJ ^^
M ••
a; z
en
01
3
(0
C
s
•»H
^4
o
• T3
O 4)
a en
•H a
f
a 01
« 2
(0
% E
12 E
3 (0
O O
en — '
;,
«.
m
••
CN
-------
_
Jj
c
o
u
-—
.
^*
en
jj
0)
e
0
01
M
CN
00
^J* ^^ ^9^
^^ ^^* ^^
O1 Cft
^, _J i
a
•O T3
C JJ C
(Q 0) (0
rH ^
(]) i— 4 CU
0) _4 0)
u -^ u
> a >
vo so
0 0
0 O 0
A m A
JJ JJ
c e
4) 4)
JJ • JJ
C >* C
o *J o
cj < o cj lyi-** u
&] -H CU ij) U z <" M-^QuQJ
.OdJ-H>Q M CJ Od»-^>
z 2 J < e s u Z2>J<
CO
o
as
O
•o J T:
m x •£ ^
vo CJ 00 vo
*» itf ^* ^
Z OS - 2
EH Z
E-* Cd » E-»
b ^ CO Cu
^^
<0 *^
o (0 >
c
<0 01 *0
(U <0 OJ
U g U
JJ JJ
Cfl 01 «J 01
O "O -^ O
iJ W i-l C U 01
icj ™'
1 1 J
^" m in
1 - -
^« ^" in en
k «. » «.
CO - - -
* CN CN CN
• ^ *» ^
CN CN CN CN
e'-
en
r-
• o\
(0 ^
JJ T3
0) C
<0
U i-l
1) vV
>i a>
(0 ia
£ >
-
0
o
>* o
CN A
JJ
e
3)
. jj .
>i C >i
jJ O •<-»
•H • O vU i-«
^H 01 ^ C71 ^
(Q tJ jC "O <0 flj
JJ >i O CJ CJ1"* U JJ
eta cj O o
33 U Z 3 J < 6
03
O
<£
O
J T3
S J= "l
CJ vO vo
»
^ m
«• »
V •
CN CN
^ ^
CN CN
-------
—
u
C
o
CJ
,_,
•
•"»
0)
i-H
J2
(0
c_.
4)
U
01
Lt
0>
IM
0)
«
__^
4) tJ
3\
-H D»
IQ 3
> *-
4J
U
0>
y-i
u
ca
1
JJ
CO
x
cn
0)
•H
o
o
a
CA
Li
4)
£
0
cn
M
r»
r*
cn
**
-,_,-
•
^
(0
4J
CO
Ll
o>
>i
(0
£
o
IT)
i— t
CA
1 a
I ^
CM
M
§
&
O
- 5
3 *
X '
M Z
as
3* oi
^.
cn
3
0)
(0
C
6
*^^
^H
O
• •n
TJ 3
O 01
ex cn
•H CU
£
Oi n
g 3
(0 Li
(0
* e
-o e
3 (0
O O
CA •=••
J
VO
^
VO
fe
m
*•
^
^>
%
»
CN
^
CN
^~.
«W
r«
cn
i^
1
(Q
^
01
0)
u
>
to
o
*
o
A
•
4J
e
01
jj .
C >i
o -*J
• O 0) -H
jj cn«H
£ fO (0 <0
O tjlt-t Ll -U
w -H a a) u
o o> -^ > o
Z 3B J < 6
T3
if)
vO
^
Z
•»
&«
(X4
^^
<0
u
•H
C
^^J
cr
Li
t-i
>
(0
0>
Li
JJ
cn
O
u cn
0) cn
JJ (0
m u
>iCJ
0 —
J
in
^
in
•.
»
^r
^
v
«
_
CN
CN
>
cn
La
o
5
II
U
Ed
D
^
M
a»
3
O
4
M
Jj
3
0
-H
IM
11
^
u
^
01
3
>-t
(0
>
jj
U
0)
U-l
<44
0)
(U
• 1
•#
(0
5
JJ
cn
o»
o
jj
TJ
di
w
cn
3
0
vi
1 1
**
(0
Ll
JJ
C
0>
0
C
0
o
-H
(0
C
•H .
§c
o
c^
U fl
1 .
M
Z -P
C
- 0)
0 0
••^ c
*» o
10 O
JJ
cn j->
0
II 0>
u
CA
-------
Table 5. The no-observed-effect concentrations (NOEC) of the
monochlorobiphenyl through hexachlorobiphenyl isoraers and
decachlorobiphenyl determined for juvenile-adult fish and early
life stages (i.e., embryos and sac fry) of fish.
Fish NOEC (ug/L)
Chlorine Number Juvenile-Adult Embryo-Sac Fry
1 50. - 80. 2. - 3.
2 12. 0.5
3 ' 2.1 0.1
4 < 0.1 - 1.5 < 0.004 - 0.06
5 0.07 0.003
6 0.01 0.001
10 0.01 0.001
46
-------
,
(0
o
(4
o
-U
o
u
04
•§
(0
(U
(0
0»
rH
<
o
4J
cn
CO
CJ
04
*™"*
cn
rH
>1
C
1
^J
^^
o
D-
U-l
o
2
<1>
o
cn
M
cn
3
0
•H
U
nj
>
y-i
O
>,
jj
•••1
o
.^
X
g
•
U3
0)
rH
.a
to
EH
a>
u
c
(U
u
(I)
144
5
<-»
0) i-3
3 \
H a>
(0 3
> -*
4J
O
(U
U-l
U-l
u
«
T3
0
£.
4J
(11
cn
0)
•H
U
s,
cn
u
0)
£
O
cn
M
cn
CO
•H
U
Ui
(0
X
»0
c
(0
*~»
0) CM
u r-
O o\
£2
• •
o r^
m v
a>
V
^^
(0
4J
a
3
O CJ
CJ ID U
*r U O
rH W Z
u
«c
3
*
3|
**
•^ts
wn"
CN
cl -
M|Z
•^
>1
4J •
•H J
C fa
i -
go
O
CJ -H
0) X
c c cu
•H O Z
rJ *J
(0 ^ >W
SCO
(Q
rH rH IM
(8 a-H
lj O 3
3 4J CJJ
.U >i
-H
a
o o o
O O O
0 0 -H
O rH A
o o o o o
o o o o o
O >H O H O
^ A rH A O
V V rH
A
O O
o o
O -H
5
o
§
o
s:
cn
cn
j
>i
z
X
cu
M
CO
g
G
h-3
CJ
M
Q
o o cn
0 CJ 0 CJ CJ J
rH M rH U « SM
CJ O CJ O O Z
•J Z J Z Z U
X
p*
1^
CO
o
(X
o
u
X
CJ
^•4
Oj
EH
O
rH^
Si
s
3
10 -H
sa
O €
4-1 (Q
o o
u
ex e
3
T3
CD
4J -H
(0 CM
•H «H
•H O
•H CJ
CJ —'
I
CN
CN CN
1
in
-------
~~~
Jj
c
o
CJ
vo
V
A
4)
«<•
^
0) J
.H o»
(0 3
> — •
4J
0
H
V V ^ fH *H fH
V A A
,^
OP
js cn cn „,
4J 0 0 0 J J 2
3OCJ OCJ OOCJ > CJ >* CJCJCJ *Z CJ
o M a ,-H cd ,H •?» H z u z auu g Q
ucjO QO cjcjO ca o M ooo jg O
a J z J z J u z z z s zzz ~ z
cu cu g
H^ ^^ LJ
§ § §
as « 2
o 9 -o
jr B x J
m CJ CJ g
fc C6 &^ s^
Z &^ Z £j
Ed cd =:
CO ^ (X =
g
3
as -i
0 >i
N Q,
O E
0 0
a E
3
1) "O
4J •«•(
(Q Qj
•^ ^H
-H O
•H CJ
CJ >-»
1
m
I
- l m *r
1 in m » *
• ^ ^ • ••
in in ^> ^r ro
I » » » «. *
I i in m ^ - •* m
vo in* % » ro •• »
^ ^» » vv» M
^» «*m «N cMf^m CN
* * * _» _» •• _• *
4J •
a c
Q) O
JJ
*y o
Lj M^
i** ^^
<»~* s
CQ -H 0
.a* -H
o ^
i g ^
1 I'M JB>
o o
C-H dP
.C O
II 2 0>
CJ C C
Cd 0 0
o -^ •*
z -u -u
ITJ (Q
^ li .^
* ** ^^
0) -U 3
3 c a
•H c
O 4J
•u o en
O HJ
0) II -U
vw
u-i O u-i
0) 0 0
^
d) CJ £
JJ J Jj
(0 S
E •> 0
•^ C U
AJ o a^
S'^-o
(0 (U
o ^ o
4J JJ 3
C T3
TJ (U 0)
35 o u
w c
3 0 s,
O CJ
ui -^
C (U £
O > 5
JJ 4J C
(0 U O
U (U -«
^J ^j ^ i
c«w
-------
Figure 1. A polychlorinated biphenyl (PCS) is a family of
compounds which consists of biphenyl that has been chlorinated at
10 possible sites. For example, monochlorobiphenyl has been
chlorinated at a single site and decachlorobiphenyl has been
chlorinated at all 10 sites. A o_,o^-Cl substituted PCB refers to
the chlorination of the 2,2',6, and 6' sites which are the ortho-
ortho prime sites on the biphenyl.
49
-------
Figure 2. Relationships between chlorine number of a PCB, n-
octanol/water pertition coefficient (Kow), and acute (96h) median
lethal concentration (LC50) for rainbow trout. The relationship
for Kow and chlorine number was modified from Wasik et al.
(1982, Fig. 3, p. 10). The LCSOs for the monochloro-biphenyl
isomers for tetrachlorobiphenyl are from Table 3. The
relationship between 96-h LC50 and chlorine number of a PCB is
defined by the regression equation: log LC50 (ug/L) = 3.16 - 0.27
Chlorine no. (R2 = 0.92; N = 4) and probably becomes asymptotic
around log Kow 6-7. Rationale is provided in Sections
III.G.l.b. and c.
50
-------
-------
Figure 3. Relationships between chlorine number of a PCB, n-
octanol/water partition coefficient (Row), and no-observed-effect
concentration (NOEC) for rainbow trout. The relationship for Kow
and chlorine number was modified from Wasik et al. (1982, Fig. 3,
p. 10). The NOECs for monochlorobiphenyl are from Section
III.A.I; the NOECs for tetrachlorobiphenyl are from Section
III.D.I. and Table 3. The relationship bewteen NOEC and chlorine
number of a PCB is defined by the regression equation: log NOEC
(ug/L) = 2.53 -0.74 chlorine no. (R2 = 0.87) and probably becomes
asymptotic around log Kow 6-7. Rationale is provided in
Sections III.G.l.b. and c.
52
-------
[T] (i/e.1) 030 N Sol
CM
I
CO
I
-------
Figure 4. Relationship between chlorine number of a PCB, n-
octanol/water partition coefficient (Row), and acute (96h) median
lethal concentration (EC50) for the aquatic invertebrates:
Gammarus pseudolimnaeus (a scud or amphipod) and Daphnia magna.
The relationship for Row and chlorine number was modified from
Wasik et al. (1982, Fig. 3, p. 10). The ECSOs for the PCB
isomers are from Table 4. The relationship between 96-h EC50 and
chlorine number for PCB isomers from monochlorobiphenyl to
tetrachlorobiphenyl (for log Row's less than 6) is defined by the
linear regression equation: log EC50 (ug/L) = 3.04 - 0.411
Chlorine no. (R2 = 0.92, N = 7). The ECSOs for the
pentachlorobiphenyl and hexachlorobiphenyl isomers are greater
than would be predicted by the regression equation (i.e., not as
toxic as tetrachlorobiphenyls). Pentachlorobiphenyl and
hexachlorobiphenyl isomers are apparently becoming to water
insoluble (log Row's of about 6 or greater) to cause acute
toxicity at concentrations lower than about 30 ug/L, which is the
EC50 for tetrachlorobiphenyl.
54
-------
[T]
-------
VIII. REFERENCES
Anonymous. 1980. The invisible menance: contaminants in the
Great Lakes. Madison, Wisconsin: the University of Wisconsin Sea
Grant Program. WIS-SG-80-133.
Aulerich RJ, Ringer RK. 1977. Current status of PCB toxicity to
mink, and effect on their reproduction. Arch Environm Contain
Toxicol 6:279-292.
Bengtsson B-E. 1980. Long-term effects of PCB (Clophen A50) on
growth, reproduction and swimming performance in the minnow,
Phoxinus phoxinus. Water Research 14:681-687.
Biggs DC, Powers CD, Rowland RG, O'Connors HB, Wurster CF.
1980. Uptake of polychlorinated biphenyls by natural
phytoplankton assemblages: field and laboratory determination of
14C-PCB particle-water index of sorption. Environ Pollut Ser A
22:101-110.
Bleavins MR, Aulerich RJ, Ringer RK. 1980. Polchlorinated
biphenyls (Aroclors 1016 and 1242): effects on survival and
reproduction in mink and ferrets. Arch Environ Contain Toxicol
9:627-635.
Branson DR, Blau GE, Alexander HC, Neely WB. 1975.
Bioconcentration of 2,2',4,4'-tetrachlorobiphenyl in rainbow
trout as measured by an accelerated test. Amer Fish Soc 104:785-
792.
Broyles RH, Noveck MI. 1979_a_. Uptake and distribution of
2,5,2',5'-tetrachlorobiphenyl in developing lake trout. Toxicol
Appl Pharmacol 50:291-298.
Broyles RH, Noveck MI. 1979b_. Uptake and distribution of
2 , 4 , 5,2',4',5'-hexachlorobiphenyl in fry of lake trout and
chinook salmon and its effects on viability. Toxicol Appl
Pharmacol 50:299-308.
Bruggeman WA, Martron LBJM, Kooiman D, Hutzinger 0. 1981.
Accumulation and elimination kinetics of di-, tri- and
tetrachlorobiphenyls by goldfish after dietary and aqueous
exposure. Chemosphere 10:811-832.
Dahlgren RB, Linder RL. 1971. Effects of polychlorinated
biphenyls on pheasant reproduction, behavior, and survival. J
Wildl Mngmt 35:315-319.
Davis TS,Pyle JL, Skillings JH, Danielson ND. 1981. Uptake of
polychlorobiphenyls present in trace amounts from dried municipal
sewage sludge through an old field ecosystem. Bull Environ
Contain Toxicol 27:689-694.
56
-------
DeFoe DL, Veith GD, Carlson RW. 1978. Effects of Aroclor 1248
and 1260 on the fathead minnow (Pimephales promelas). J Fish Res
Board Can 35:997-1002.
Dill DC, Mayes MA, Mendoza CG, Boggs GU, Emraitte JA. 1982.
Comparison of the toxicities of biphenyl, monochlorobiphenyl, and
2,2',4,4'-tetrachlorobiphenyl to fish and daphnids. In: Aquatic
toxicology and hazard assessment: fifth conference. Pearson JG,
Foster RB, Bishop WE, eds. Philadelphia, PA: American Society
for Testing and Materials. ASTM STP 766.
Dive D, Erb F, Leclerc H, Priem M, Colein M. 1976. Toxicity and
bioacumulation of polychlorinated bihphenyl isomers in the
ciliated protozoa Colpidium campylum (Stokes). Eur J Toxicol
9:105-111. (in French; English trans.)
Freeman HC, Sangalang GB. 1977. A study of the effects of
methyl mercury, cadmium, arsenic, selenium, and a PCB (Aroclor
1254) on adrenal and testicular steroidogenesis in vitro, by the
gray seal Halichoerus grypus Arch Environm Contam Toxicol 5:369-
383.
Gilbertson M, Hale R. 1974. Characteristics of the breeding
failure of a colony of Herring Gulls on Lake Ontario. Canadian
Field Naturalist 88:356-358.
Glooschenko V, Glooschenko W. 1975. Effect of polychlorinated
biphenyl compounds on growth of Great Lakes phytoplankton. Can J
Bot 53:653-659.
Goldstein JA, Hickman P, Bergman H, McKinney J, Walker MP.
1977. Separation of pure polychlorinated bipheny isomers into
two types of inducers on the basis of induction of cytochrome P-
450 or P-448. Chem Biol Interact 17:69-87.
Heinz GH, Katsma DE, Swineford DM. Unpublished manuscript.
Significance of high PCB residues in birds from the Sheboygan
River, Wisconsin, 1976-80. Laurel, Maryland: Patuxent Wildlife
Res Center, US Fish and Wildlife Service.
Helle E, Olsson M, Jensen S. 1976. PCB levels correlated with
pathological changes in seal uteri. Ambio 5:261-263.
Henny CJ, Blus LJ, Gregory SV, Stafford CJ. 1981. PCBs and
organochlorine pesticides in wild mink and river otters from
Oregon, In: Worldwide Furbearer Conference Proceedings. Chapman
JA, Pursely D, eds. Frostburg, Maryland: Worldwide Furbearer
Conference, Inc.
Hermens J. 1982. QSAR's in aquatic toxicology studies. Fourth
European Symposium on Chemical Structure - Biological Activity
Quantitive Approaches, Bath, England, September 6-9, 1982.
Abstract.
57
-------
Jensen S. , Kihlstrcm JE, Olsson M, Lundberg C, Orberg J, 1977.
Effects of PCB and DDT on mink (Mustela vision during the
reproductive season. Ambio 6(4):239.
Keil JE, Priester LE, Sandifer SH. 1971. Polychlorinated
biphenyl (Aroclor 1242): effects of uptake on growth, nucleic
acids and chlorophyll of a marine diatom. Bull Environm Contain
Toxicol 6:156-159.
Konemann H. 1981. Quantitative structure-activity relationships
in fish toxicity studies. Part 1: relationships for 50
industrial pollutants. Toxicology 19:209-221.
Kricher JC, Bayer CL. 1977. Depression of primary productivity
of Aroclor 1232 in an interspecific lentic algal assembledge.
Bull Environ Contam Toxicol 18:14-17.
Lincer JL, Peakall DB. 1973. PCB pharmacodynamics in the ring-
dove and early gas chromatographic peak diminution, Environ
Pollut 4:59-68.
Mauck WL, Mehrle PM, Mayer FL. 1978. Effects of the
polychlorinated biphenyl Aroclor 1254 on growth, survival, and
bone development to brook trout (Salvelinus fontinalis). J Fish
Res Board Can 35:1084-1088.
Mayer FL, Mehrle PM, Sanders HO. 1977. Residue dynamics and
biological effects of polychlorinated biphenyls in aquatic
organisms. Arch Environm Contam Toxicol 5: 501-511.
Moore Jr SA, Harriss RC. 1972. Effects of polychlorinated
biphenyl on marine phytoplankton communities. Nature 240:356-
358.
National Research Council. 1979. Polychlorinated biphenyls.
Washington, DC: National Academy of Sciences.
Nebecker AV, Puglisi FA. 1974. Effect of polychlorinated
biphenyls (PCBs) on survival and reproduction of Daphnia,
Gammarus, and Tanytarsus. Trans Amer Fish Soc 103:722-728.
Nebecker AV, Puglisi FA, DeFoe DL. 1974. Effect of
polychlorinated biphenyl compounds on survival and reproduction
of the fathead minnow and flagfish. Trans Amer Fish Soc 103:562-
568.
Nowicki HG, Norman AW. 1972. Enhanced hepatic metabolism of
testosterone, 4-androstene-3,17-dione, and estradiol-17B in
chickens pretreated with DDT or PCB. Steroids 19:85-99.
O'Conners AJ, Mahanty HK. 1979. Growth responses of unicellular
algae to polychlorinated biphenyls: new evidence for
photosynthetic inhibition. Mauri Ora 7:3-17.
58
-------
O'Conners HB Jrr Wurster CF, Powers DC, Biggs DC, Rowland RG.
1978. Polychlorinated biphenyls may alter marine trophic
pathways by reducing phytoplankton size and production. Science
201:733-739.
Peakall DB, Lincer JL, Bloom SE. 1972. Embryonic mortality and
chromosomal alterations caused by Aroclor 1254 in ring doves.
Environ Health Perspectives 1:103-104.
Peterson RE, Guiney PD. 1971. Disposition of polychlorinated
biphenyls in fish. IN: Pesticides and xenobiotic metabolism in
aquatic organisms. American Chemical Society. ACS Symp Ser
99:21-36.
Pizza JC, O'Connor JM. 1983. PCB dynamics in Hudson River
striped bass. II. Accumulation from dietary sources. Aquatic
Toxicology 3:313-327.
Poland A, Glover E. 1977. Chlorinated biphenyl induction of
aryl hydrocarbon hydroxylase activity: a study of the structure-
activity relationship. Mol Pharmacol 13:924-938.
Reijnders PJH. 1980. Organochlorine and heavy metal residues in
harbour seals from the Wadden Sea and their possible effects on
reproduction. Netherlands J Sea Res 14(l):30-65.
Risebrough RW, Rieche P, Peakall DB, Herman SG, Kirven MN.
1968. Polychlorinated biphenyls in the global ecosystem. Nature
220:1098-1101.
Rubinstein NI, Lores E, Gregory NR. 1983. Accumulation of PCBs,
mercury and cadmium by Nereis virens, Mercenaria mercenaria and
Palaemonetes pugio from contaminated harbor sediments. Aquatic
Toxicology 3:249-260.
Sangalang GB, Freeman HC, Crowell R. 1981. Testicular
abnormalities in cod (Gadus morhua) fed Arochlor 1254. Arch
Environ Contain Toxicol 10:617-626.
Schimmel SC, Hansen DJ, J Forester. 1974. Effects of Aroclor
1254 on laboratory-reared embryos and fry of sheepshead minnows
(Cyprinodon variegatus). Trans Amer Fish Soc 103:582-586.
Schmitt CJ, Ludke JL, Walsh DF. 1981. Organochlorine residues
in fish: national pesticide monitoring program, 1970-74. Pest
Monit J 14:136-206.
Seelye JG, Mac MJ. 1981. Size-specific mortality in fry of lake
trout (Salvelinus namaycush) from Lake Michigan. Bull Environ
Contain Toxicol 27:376-379.
Shaw GR, Connell DW. 1980j_. Polychlorinated biphenyls in the
Brisbane River Estuary, Australia. Mar Poll Bull 11:356-358.
59
-------
Shaw GR, Connell DW. 1980J3. Relationships between steric
factors and bioconcentration of polychlorinated biphenyls (PCB's)
by the sea mullet (Mugil cephalus Linnaeus). Chemosphere 9:731-
743.
Skea JC, Simonin HA, Dean HJ, Colquhoun JR, Spagnoli JJ, Veith
GD. 1979. Bioaccumulation of Aroclor 1016 in Hudson River
Fish. Bull Environm Contain Toxicol 22:332-336.
Spigarelli SA, Thommes MM, Prepejchal W. 1983. Thermal and
metabolic factors affecting PCB uptake by adult brown trout.
Environ Sci Technol 17:88-94.
Stalling DL, Huckins JN, Petty JD, Johnson JL, Sanders HO.
1979. An expanded approach to the study and measurement of PCBs
and selected planar halogenated aromatic environmental
pollutants. Ann NY Acad Sci 320:48-59.
Sugiura K, Ito N, Matsumoto N, Mihara Y, Murata K, Tsukakoshi Y,
Goto M. 1978. Accumulation of polychlorinated biphenyls and
polybromonated biphenyls in fish: limitations of correlation
between partition coefficients and accumulation factors.
Chemosphere 9:731-736.
Swain WR. 1980. An ecosystem approach to the toxicology of
residue forming xenobiotic organic substances in the Great
Lakes. A manuscript invited by the Environmental Studies Board
of the National Research Council, National Academy of Science.
Grosse lie, Michigan: Large Lakes Research Station, US
Environmental Protection Agency.
Swain WR. 1982. What is the scientific basis for the present
concern in the Great Lakes? Presented before the International
Symposium on PCBs in the Great Lakes. March 15-17, 1982 at
Michigan State University, East Lansing, Michigan.
Thomann RV. 1978. A size dependent model of hazardous
substances in aquatic food chain. Duluth, Minnesota:
Environmental Research Laboratory, US Environmental Protection
Agency. EPA-600/3-78-036. Available from NTIS, Springfield,
VA. PB 281-009.
United States Environmental Protection Agency (USEPA). 1977.
Toxic Pollutant Effluent Standards: Standards for
polychlorinated biphenyls (PCBs); final decision. Federal
Register 42:6532-6555.
United States Environmental Protection Agency (USEPA). 1980.
Ambient Water Quality Criteria for polychlorinated biphenyls.
Washington, DC: Office of Water Regulations and Standards,
USEPA. EPA 440/5-80-068. Available from NTIS, Springfield,
VA. PB 81-117798.
60
-------
Versar Inc. 1983. Exposure assessment for incidentally produced
polychlorinated biphenyls (PCBs), Vol II. Preliminary draft
report. Task 21. Washington, DC: Exposure Evaluation Division,
Office of Toxic Substances, US Environmental Protection Agency.
Contract 68-01-6271.
Von Westerhagen H, Rosenthal H, Dethlefsen V, Ernst W, Harms U,
Hansen P-D. 1981. Bioaccumulating substances and reproductive
success in baltic flounder Platichthys flesus. Aquat Toxicol
1:85-99.
Vreeland V. 1974. Uptake of chlorobiphenyls by oysters.
Environ Pollut 6:135-140.
Wasik SP, Tewari YB, Miller MM, Martire DE, Ghodbane S. 1982.
Final report on project "Water solubility and octanol/water
partition coefficient of polychlorinated biphenyls and other
selected substances - Task 1C." Draft final report. National
Bureau of Standards. Washington, DC: Office of Pesticides and
Toxic Substances, U.S. Environmental Protection Agency.
Interagency Agreement EPA-80-D-X0958.
Weininger D. 1978. Accumulation of PCBs by lake trout in Lake
Michigan. Ph.D. dissertation, Madison, WI: Univ of Wisconsin.
Zimmerman N. 1982. Polychlorinated biphenyls in Great Lakes
fish. Lansing, Michigan: Toxic Substance Control Commission, 815
Washington Square Building 48909.
U S. Environmental Protection Agency
Region V, Library
230 South Dearborn Street
Chicago, Illinois 60604
61
------- |