EPA-822-R-93-007
  vvEPA
              United States
              Environmental Protection
              Agency
             Office of Water (WH-586)    EPA-822-R-93-O07
             Office of Science and Technology April 1993
             Washington, DC 20460
Great Lakes
Water Quality Initiative
Criteria Documents for
the Protection of Wildlife
(PROPOSED)
                              ENVIRONMENTAL
                                PROTECTION
                                AGf-IMCV
              DDT
              Mercury
              2,3,7,8-TCDD
              PCBs

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                                      EPA-822-R-93-007
                 AGENCY                     April 1993

               DALLAS, TEXAS
    Great Lakes Water Quality Initiative
  Criteria Documents for the Protection of
          Wildlife (PROPOSED)
    DDT; Mercury; 2,3,7,8-TCDD; PCBs
      Office of Science and Technology
             Office of Water
United States Environmental Protection Agency
         Washington, D.C. 20460

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                             ACKNOWLEDGEMENTS
Authors

Steven Bradbury
U.S. EPA, Environmental Research Laboratory
Duluth, MN

Cynthia Nolt
U.S. EPA, Office of Science and Techno'^gy,
Office of Water, Washington, DC

Beth Goodman
Wisconsin Department of Natural Resources
Madison, WI
Kenneth Stromborg
U.S. Fish and Wildlife Service
Green Bay, WI

John Sullivan
Wisconsin Department of Natural Resources
Madison, WI
Technical Support
Patrick Fitzsimmons
ASCI, Duluth, MN
Elizabeth Mooney
SAIC, Fairfax, VA
Isaac Diwan
SAIC, Fairfax, VA
Jeffrey Peterson
ManTech Environmental
Corvallis, OR

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         proved for publication as a support document for the  Great Lakes Water Quality
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     This document Is available for a fee upon written request or telephone call to:

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 CHAPTER 3
 Tier  I Wildlife Criteria for
 2,3,7,8-Tetrachiorodibenzo-p-dioxin (2,3,7,8-TCDD)	

 I.   Literature Review	   3-1

 D.   Calculation of Mammalian Wildlife Value	   3-1
     i.   Acute Toxicity	   3-1
     ii.  Chronic Toxicity	   3-2
     iii. Mammalian Wildlife Value Calculation	   3-4
     iv. Sensitivity Analysis for Mammalian Wildlife Value  	   3-5

 m.  Calculation of Avian Wildlife Value  	   3-6
     i.   Acute Toxicity	   3-6
     ii.  Chronic Toxicity	   3-6
     iii. Avian Wildlife Value Calculation  	   3-7
     iv. Sensitivity Analysis of Avian Wildlife Value  	   3-8

 IV.  Great Lakes Wildlife Criterion	   3-9
     i.   Discussion of Uncertainties  	   3-9

 V.   References   	    3-10
CHAPTER 4
Tier I Wildlife  Criteria for  Polychlorinated Biphenyls  (PCBs)	

I.   Literature Review	    4-1

H.  Calculation of Mammalian Wildlife Value	    4-1
    i.   Acute Toxicity	    4-1
    ii.  Chronic Toxicity	    4-1
    iii.  Mammalian  Wildlife Value Calculation	    4-5
    iv.  Sensitivity Analysis for Mammalian Wildlife Value	    4-7

HI. Calculation of Avian Wildlife Value  	    4-7
    i.   Acute Toxicity	    4-7
    ii.  Chronic Toxicity	    4-8
    iii.  Avian Wildlife Value Calculation  	    4-11
    iv.  Sensitivity Analysis of Avian Wildlife Value  	    4-12

IV. Great Lakes Wildlife Criterion	    4-13
    i.   Discussion of Uncertainties 	    4-13

V.  References  	    4-14

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 Contents
CHAPTER 1
Tier I Wildlife Criteria forp,p'- Dichlorodiphenyltrichloroethane
(DDT) and Metabolites

I.   Literature Review	   1-1

n.  Calculation of Mammalian Wildlife Value	   1-1
    i.   Acute Toxicity Studies	   1-1
    ii.  Chronic Toxicity Studies	   1-2
    iii.  Mammalian Wildlife Value Calculation	   1-3
    iv.  Sensitivity Analysis for Mammalian Wildlife Value  	   1-4

m.  Calculation of Avian Wildlife Value  	   1-5
    i.   Acute Toxicity Studies	   1-5
    ii.  Chronic Toxicity Studies	   1-6
    iii.  Avian Wildlife Value Calculation  	   1-7
    iv.  Sensitivity Analysis for Avian Wildlife Value	   1-9

IV.  Great Lakes Wildlife Criterion	   1-9
    i.   Discussion of Uncertainties  	   1-10

V.  References 	   1-10
CHAPTER 2
Tier I Wildlife Criteria for Mercury (Including Methylmercury)

I.   Literature Review	   2-1

II.  Calculation of Mammalian Wildlife Value	   2-1
    i.   Acute Toxicity Studies	   2-1
    ii.  Chronic Toxicity Studies	   2-1
    iii. Mammalian Wildlife Value Calculation	   2-2
    iv.  Sensitivity Analysis for Mammalian Wildlife Value  	   2-3

m.  Calculation of Avian Wildlife Value  	   2-4
    i.   Acute Toxicity Studies	   2-4
    ii.  Chronic Toxicity Studies	   2-5
    iii. Avian Wildlife Value Calculation  	   2-7
    iv  Sensitivity Analysis for Avian Wildlife Value	   2-8

IV.  Great Lakes Wildlife Criterion	   2-9
    i.   Discussion of Uncertainties  	   2-9

V.  References 	".	   2-9

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CHAPTER 1
Tier I Wildlife  Criteria  forp,p'-
Dichlorodiphenyltrichloroethane  (DDT)
and Metabolites
                             Contents


I.  Literature Review	  1-1

n.  Calculation of Mammalian Wildlife Value	  1-1

   i.  Acute Toxicity Studies	  1-1

   ii.  Chronic Toxicity Studies .	  1-2

   iii. Mammalian Wildlife Value Calculation	  1-3

   iv. Sensitivity Analysis for Mammalian Wildlife Value  	  1-4

El. Calculation of Avian Wildlife Value 	  1-5

   i.  Acute Toxicity Studies	  1-5

   ii.  Chronic Toxicity Studies	  1-6

   iii. Avian Wildlife Value Calculation 	  1-7

   iv. Sensitivity Analysis for Avian Wildlife Value	  1-9

IV. Great Lakes Wildlife Criterion	  1-9

   i.  Discussion of Uncertainties  	  1-10

V.  References  	  1-10

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Tier  I  Wildlife  Criteria  forp,p'-
Dichlorodiphenyltrichloroethane
(DDT) and  Metabolites
I. Literature Review
      A review of mammalian and avian toxicity data for p,p '-dichlorodiphenyl-trichloroethane
   (DDT) and its metabolites was based on literature received through computer-based (CAS and
   BIOSES) as well as manual searches. A total of 36 references were screened for dose-response
   data. The majority of those references consisted of studies on avian species. Those references
   which  were reviewed in detail, specifically those that contain dose-response data, are cited in
   Section V.


II. Calculation of Mammalian Wildlife Value
   /. Acute Toxicity Studies
      According to the RTECS database (NIOSH, 1992), the oral LDjo values for DDT range
   from 87 mg/kg for the rat to more than 5000 mg/kg for the hamster (See Table 1-1). LDX values
   for DDT from other exposure routes range from 0.91-1931 mg/kg (NIOSH, 1992).

                   Table 1-1. Mammalian Acute Toxicity Values
Route
oral
oral
oral
oral
oral
oral
oral
oral
oral
dermal
dermal
dermal
i.p.
i.p.
Species
rat
rat
mouse
dog
monkey
cat
rabbit
guinea pig
hamster
rat
rabbit
guinea pig
rat
mouse
LDM (mg/kg)
87
152.3'
135
150
200
250
250
150
> 5000
1931
300
1000
0.91
32
                                                                    1-1

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                      Table 1-1. Mammalian Acute Toxicity Values (Cont.)
Route
s.c.
s.c.
s.c.
i.v.
i.v.
i.v.
i.v.
i.v.
i.v.
unreported
unreported
Species
rat
rabbit
guinea pig
rat
mouse
dog
monkey
cat
rabbit
rat
mammal
LDM (mg/kg)
1500
250
900
68
6.85
150
50
40
50
300
200
                     Source: NIOSH (1992), except for' Mijavila et al. (1981).

    ii. Chronic Tox/city Studies
         No suitable subchronic or chronic studies were found for mammalian wildlife in which dose-
    response data was reported. Gilbert (1969) did examine the toxicity of DDE to mink, although no
    dose-response data could be developed because exposures to DDT were intermittent and total
    DDT intake was unquantifiable, due to the experimental protocol that was used. Gilbert (1969)
    fed 10 male and 10 female mink a contaminated fish ration containing 0.58 ppm DDE. Three
    male and 2 female mink died within 20 days. The remainder of the experimental group was then
    maintained on a control ration, and intermittently on contaminated feed, for two different periods
    lasting up to 47 days. DDE residues were found to be greatest in  the liver and brain tissues of the
    experimental animals and the spleen, adrenal glands, and testes were heavier in experimental
    animals than in  controls. The whelping rate among  the experimental animals was approximately
    half that of controls, and the average number of live kits 24 hours after birth was significantly
    reduced among the experimental females. Average in utero loss and average total loss of kits was
    also greater among the experimental group than among the controls.
         Chronic and subchronic studies of the toxicity of DDT to mammals have been conducted
    using typical laboratory animals. In one subchronic study (Mitjavila et al., 1981), OFA Sprague
    Dawley rats (male, 32 per group) were administered p,p' DDT in an oil vehicle by gavage at a
    dose of 14.5 mg/kg/day for up to 52 days. Liver weight in the treated group was 20 percent
    greater than in the control group due to cellular hypertrophy induced by the DDT, and the level
    of total lipids was 30 percent less in the treated group than in the  controls.
         In another subchronic rat-feeding study (Laug et al., 1950),  weanling rats (15/sex/group)
    were fed  commercial-grade DDT (81 percent p,p'-  DDT, 19 percent o,p -DDT) at levels of 0, 1,
    5, 10 or 50 ppm for 1-27 weeks. The critical effect was liver toxicity, demonstrated  as relatively
    mild dose-dependent histopathologic changes in hepatocytes at doses of 5 ppm or higher. These
    changes included hepatocellular hypertrophy, increased cytoplasmic oxyphillia, and peripheral
    basophilic cytoplasmic granules. The NOAEL was  1 ppm.  Based  on a rat body weight of 0.20 kg
    and food  ingestion rate of 0.01 kg/d (i.e., 5 percent body weight per day) (NIOSH, 1992), the
    LOAEL for liver effects war 0.25  mg/kg/day (5 ppm) and  the NOAEL for liver effects was 0.05
    mg/kg/day.
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     Both the studies by Mitjavila et al. (1981) and Laug et al. (1950) would not be acceptable
for derivation of a mammalian wildlife value (WV) because of study duration, requirement for
dose-response data and/or because the toxicity endpoint is not an acceptable endpoint as defined in
the wildlife criteria methodology provided in Appendix D to 40 CFR 132.  (For the purpose of
wildlife criteria derivation, an acceptable subchronic or chronic endpoint is one that affects
organismal growth or viability, or reproductive or developmental success or any other endpoint
which is, or is directly related to, parameters that influence population dynamics.) These studies
are presented here to provide relative perspective for doses at which lexicological impacts occur.
     In a 2-year reproduction study (Fitzhugh, 1948), rats were provided a diet that contained 0,
10, SO,  100, and 600 ppm DDT. The number of litters, number of live young at birth, average
weight at birth, and the number of young surviving through the weaning period were quantified.
The number of litters, number of living young at birth, and average weight at birth did not appear
to differ with dosage level. At a concentration of 50 ppm DDT, the number of weanling rats was
reduced by approximately 20 percent. The NOAEL was 10 ppm DDT since no effect was
observed at that level. Based on a rat body weight of 0.20 kg  and a food ingestion rate of 0.01
kg/d (NIOSH, 1992), the LOAEL derived from this study was 2.5  mg/kg/day (50 ppm) and the
NOAEL derived for calculation of a mammalian wildlife value was 0.50 mg/kg/day.

The results of the mammalian studies described above are summarized in Table 1-2.
                   Table 1-2. Summary of Chronic Mammalian Studies
Species
Mink
Rat
Rat
Rat
LOAEL
(mg/kg/day)
n/a
14.5
0.25
2.5
NOAEL
(mg/kg/day)
n/a

0.05
0.5
Toxic Effect Observed
Reproductive
Liver toxicity
Liver toxicity
Reproductive
Reference
Gilbert, 1969
Mitjavila et al.,
1981
Laug et al., 1950
Fitzhugh et al.,
1948
     The study by Fitzhugh (1948) was selected for developing Tier I mammalian wildlife values
because the Fitzhugh (1948) study consists of repeated oral exposures for over a 90-day period,
and reports observed reproductive effects from chronic exposures. Therefore, this study fulfills
the requirements for an appropriate study for wildlife criteria development as described in
Appendix D to 40 CFR 132. The LOAEL for reproductive effects reported in Fitzhugh (1948)
was 2.5 mg/kg/day (50 ppm) and the NOAEL was 0.5 mg/kg/day (10 ppm).

///. Mammalian Wildlife  Value Calculation
     In calculating a Tier I wildlife value, a species sensitivity factor (SSF) of 1 to 0.01 is
recommended  in Appendix D to 40 CFR 132 to accommodate differences in interspecies toxicity.
Because of the paucity of subchronic or chronic mammalian toxicity studies assessing the toxicity
of DDT or its  metabolites, a SSF of 0.1 is used to reflect the uncertainty in extrapolating toxicity
data from the rat to the mink and river otter.
     DDT bioaccumulation factor  (BAF) values for Trophic Level 3 and Trophic Level 4 were
derived based on the Great Lakes Water Quality Initiative procedure to determine a
bioaccumulation factor presented in Appendix B to 40 CFR  132.
                                                                                       1-3

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    Input parameters for the wildlife equation are presented below.
                      NOAEL (mammalian)
                    BAF3 (Trophic Level 3)
                    BAF4 (Trophic Level 4)
                                     SSF
                            0.5 mg/kg body weight/day
                            1,000,000 f/kg body weight
                            3,000,000 //kg body weight
                            0.1 (mink and otter)
    Body weights (WtJ, ingestion rates (FJ, and drinking rates (W^ for mink and river otter are
    presented in Table D-2 of Appendix D to 40 CFR 132 and shown below.
                                  WtA (mink)
                                  WtA (otter)
                                    FA (mink)
                                    FA (otter)
                                   WA (mink)
                                   WA (otter)
                                1.0kg
                                8.0kg
                                0.15 kg/day
                                0.9 kg/day
                                0.099 f/day
                                0.64 f/day
    The wildlife equations and calculations of mammalian wildlife values are presented below:
     WV (mink)

     WV (mink)

     WV (mink)

     WV (otter)

     WV (otter)

     WV (otter)
                          [NOAEL x SSF] x WtA(mrt0
                       WAtlMlk) + [(1.0)(FA(BWlklxBAF3)]
               	(0.5 mg/kg/d x 0.1) 1.0kg	
                0.099 ltd + [(1.0K.15 kg/d x 1,000,000 //kg)]
                                 333 pg//
             	[NOAEL x SSF) x WtA(ong,	
             WAlonirl + H0.5HFA(od_, x BAF3)  + (0.5)(FA(ottlfl x BAF4)].
                         (0.5 mg/kg/d xO.1) 8.0kg
0.64 lid + [(0.5K0.90 kg/d x 1,000,000 I/kg) +  (0.5)(0.90 kg/d x 3,000,000 */kg)J
                                 222 pg/l
    The geometric mean of these two mammalian wildlife values results in
                           WV (mammalian)   = e
                           WV (mammalian)   = e
                                                 (llnWV(mink) + In WVIotw W2I
                             (lln 333 pg/f + In 222 PQ//I/2)
                           WV (mammalian)   = 270 pg/f.
    iv.  Sensitivity Analysis for Mammalian  Wildlife  Value
         The values of the various parameters used to derive the mammalian wildlife value presented
    above represent the most reasonable assumptions. The purpose of this section is to illustrate the
    significance of thest assumptions and the variability in the mammalian wildlife value if other
    assumptions are made for the values of the various parameters from which the mammalian
    wildlife value is derived. The intent of this section  is to let the  risk manager know, as much as
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    possible, the influence on the magnitude of the mammalian wildlife value of the assumptions
    made in its derivation.
        In developing a mammalian wildlife value for DDT, reproductive effects of toxicity were
    judged to be of importance to protect wildlife populations. Toxic effects of DDT on the liver
    were observed at concentrations lower than the NOAEL for reproductive effects. The NOAEL
    determined in the Laug et al. (1950) rat study suggests that toxic effects on the liver can occur at
    DDT doses that are as much as 10 times lower than the NOAEL for reproduction of 0.5
    mg/kg/day determined by Fitzhugh (1948)  and used to calculate the mammalian wildlife value.
    Although the methodology does not allow for basing a wildlife criterion on a liver toxicity
    endpoint, if one used the NOAEL from the Laug et al. (1950) study, the mammalian wildlife
    value would be 27 pg/l rather man 270 pg/l.
        In estimating the hazards of DDT to mammalian wildlife, a SSF of 0.1 was used to reflect
    the uncertainty in extrapolating toxicity data from the rat to mink and river otter. Based on the
    lack of mammalian chronic toxicity data, the use of such a factor seems reasonable. Oral acute
    toxicity values (Table 1-1) show a wide variability in sensitivity within a species, but they also
    show mat the rat is among the most sensitive of the mammalian species tested for the acute effects
    of DDT. This may justify the use of an alternate and less conservative  SSF of 0.3-the geometric
    mean of 1  and 0.1. If an  intermediate SSF  of 0.3 were used with the NOAEL determined in the
    Fitzhugh (1948) study, the mammalian wildlife value would be 820 pg/l instead of 270 pg/l.
        In deriving the DDT mammalian wildlife value, it was assumed that 100 percent of the mink
    diet was comprised of fish, although this may  not necessarily be the case. This assumption may
    lead to an overestimate of DDT exposure for mink that are not primarily feeding on fish and
    aquatic invertebrates. As  indicated in the Technical Support Document  for Wildlife Criteria
    (Appendix to the Preamble  to 40 CFR 132), the fish content of a mink  diet can vary from less
    than 50 percent to the 100 percent assumed in the mink wildlife value derivation presented above.
    If it were assumed only 50 percent of a mink's diet was from aquatic resources and the remaining
    50 percent of the diet was uncontaminated, the estimated DDT exposure would be reduced by a
    factor of 2. The resulting wildlife value for the mink would be 670 pg/l, and the mammalian
    wildlife value would be 380 pg/l, rather than the mammalian wildlife value of 270 pg/l.


III. Calculation of Avian Wildlife Value
    /.  Acute Toxicity Studies
        Long-term exposure of birds to DDT  has been demonstrated to result in eggshell thinning in
    several species; however, the acute toxicity of DDT has not been well established. The RTECS
    database (NIOSH, 1992)  listed the oral LD value for chickens (Callus) as 300 mg/kg. Bernard
    (1963) observed tremors  within 7 days in robins (Turdus migratorius) ingesting feed contaminated
    with 300 mg/kg DDT. Stickel et al. (1966) reported the oral DDT LQo for bald eagles (Haliaetus
    leucocephalus) as 80 ppm following dietary exposure for 3-4 months. For the clapper rails (Rallus
    longirostris), the DDT oral LCjo value was 1612 ppm for males and 1896 ppm for females. The
    LCjo value for juvenile (2 to 3 weeks old) ring-necked pheasant (Phasianus colchicus) was 311
    ppm, while the value for  juvenile mallard ducks  (Anas platyrhynchos) was found to be  1869 ppm
    (Van Veltzen and Kreitzer,  1975).
        LCjo values for DDT concentrations in brain tissue have also been determined for avian
    species. The geometric mean brain DDT residue LCy, values ranged from 23 ppm wet weight for
    the blue jay (Cyanocitta cristata) to 109 ppm wet weight for the cardinal (Richmondena
    cardinalis) (Van Veltzen  and Kreitzer, 1975). Stickel et al. (1984) established that 300-400 ppm
    DDE wet weight in brain tissue caused death in grackles (Quiscalus guiscula), red-winged
    blackbirds (Agelaius phoeniceus), brown-headed  cowbirds (Molothrus atef) and starlings (Sturnus
    vulgaris). DDE residues in brains of two kestrels (Falco sparverius) that died following  14
    months of exposure to 2.8 ppm dietary DDE (wet weight, or 10 ppm dry weight) were 212.5 and
    301.1 ppm wet weight (Porter and Wiemeyer,  1972).
                                                                                         7-5

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    //. Chronic Toxicity Studies
         The toxicity of DDT has been documented in a number of avian species including the
    mallard (Kolaja. 1977; Heath et al., 1969; Davison and Sell, 1974), kestrels (Peakall et aJ.,
    1973), and brown pelicans (Pelecanus occidentalis) (Anderson et al., 1975).
         Numerous studies of DDT and/or DDE ingestion by mallard ducks at levels ranging from 10
    to 40 ppm in feed for a period ranging from 5 weeks prior to egg laying and through two years
    have demonstrated significant reduction in eggshell thickness (Haegele and Hudson, 1974;
    Longcore and Samson, 1973; Davison and Sell, 1973; Risebrough and Anderson, 1975; Kolaja
    and Hinton, 1977).
         Davison and Sell (1974) exposed female mallards to technical grade DDT and pure/?,/?'-
    DDT at 0, 2,  20, and 200 ppm in the diet and assessed effects on eggshell thickness. Significant
    reduction in eggshell thickness was observed at 20 ppm (the LOAEL), and the NOAEL was 2
    ppm for eggshell thickness. Lethality was observed at 200 ppm dietary DDT. Using a mallard
    body weight of 1 kg (Delnicki and Reinecke, 1986), and a feeding rate of 0.0582 kg/d determined
    using the allometric relationship provided in Appendix D to 40 CFR  132, a LOAEL value of 1.16
    mg/kg/day (20 ppm) and a NOAEL of 0.116 mg/kg/day (2 ppm) can be estimated for effects on
    eggshell thickness.
         Kolaja (1977) quantified effects of dietary DDT and DDE on mallard duck eggshell
    thickness and  weight.  Birds were exposed to dietary DDT and DDE at 0, 10 and 50 ppm.
    Eggshell thickness and weight were significantly reduced at both dose levels for either DDT or
    DDE. Using the mallard body weight and ingestion rate presented above, the LOAEL determined
    in this study is 2.91 mg/kg/day for eggshell thickness and weight.
         Heath et al. (1969) exposed mallard ducks to dietary DDT, DDE, and DDD for 2 years and
    assessed reproductive success and eggshell thinning. Ducks were exposed to dietary DDE and
    DDD in commercial feed at 10 and 40 ppm and DDT at 2.5, 10, and 40 or 25 ppm (the higher
    concentration  was reduced after breeders died). Endpoints evaluated were percent cracked eggs,
    embryo mortality, hatchling survivability, and  number of ducklings per hen. DDE severely
    impaired reproductive success at both dose levels, and duckling production per hen was reduced
    by 50 to 75 percent. The DDE LOAEL for reproductive success obtained from this study was 10
    ppm, or 0.58  mg/kg/day calculated using the default body weight and feed ingestion rate
    presented previously.  Heath et al. (1969) also reported that DDD  impaired reproductive success,
    but less severely than did DDE. DDT in the diet at concentrations of 2.5 and 10 ppm did not
    have measurable effects on reproduction. Therefore, the LOAEL for DDT in the diet of mallard
    ducks based on reproductive success is 1.45 mg/kg/day (25 ppm) and the NOAEL is 0.58
    mg/kg/day (10 ppm).
         Peakall et al. (1973) exposed American kestrels to 3, 6, and 10 ppm DDE in the diet and
    measured eggshell thickness, breaking strength, and permeability. Significant effects on each of
    these endpoints were observed at the lowest dietary concentration. Using a default kestrel body
    weight of 100 g (Bloom, 1973), and an  ingestion rate derived from the allometric relationship
    presented in the Appendix D to 40 CFR 132, the LOAEL determined for DDE in this study is   -
    0.39 mg/kg/day (3 ppm).
         Alsop (1972) compared red-winged blackbird eggs collected  during two successive seasons
    in an area ranging from Tennessee to Florida to pre-DDT eggs from  museum collections. Post-
    DDT eggs were significantly thinner than those in the museum collections. However, since no
    measurements of DDT levels were performed on either the field sampled eggs or the museum
    collection, attributing the effect to DDT alone is speculative.
         Anderson et al. (1975) studied the reproductive  success of brown pelicans off the coast of
    southern California for the years of 1969 through 1974. Concentrations of DDT and metabolites
    in anchovies, the major food source of this pelican colony, and pelican eggs were also measured
    during the course of this investigation. Over the five years, combined concentrations of DDT,
    DDD, and DDE in the food source steadily declined  from 4.27 ppm  (wet weight) in 1969 to 0.15
    ppm in 1974.  At 0.15 ppm total DDT and metabolites in the food source, the fledgling rate was
1-6

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30 percent below the estimated rate necessary to maintain a stationary population. Based on the
results of this study, a LOAEL of 0.15 ppm total DDT can be inferred for reproductive success.
Using a pelican body weight of 3.5 kg (Dunning, 1984), and the allometric equations presented in
Appendix D to 40 CFR 132, the calculated food ingestion rate for pelicans is 0.131 kg/day (dry
weight).  Since the DDT bioaccumulation factor for the pelican's food source is provided in terms
of wet weight, the calculated dry weight food ingestion rate is converted to a wet weight food
ingestion rate by multiplying by 5 (U.S. EPA, 1980). This results in an intake rate of 0.66 kg/d.
Multiplying the LOAEL (0. IS ppm) by the food ingestion rate and dividing by the pelican body
weight gives  > LOAEL of 2.82 x 10*2 mg/kg/day for reproductive success. The results of the
studies described  above are summarized in Table 1-3.
     The Anderson et al. (1975) study with brown pelicans was judged most appropriate for avian
wildlife value development because it consists of a peer-reviewed field study of a wildlife species
that provides a chemical-specific dose-response curve for reproductive success. According to the
methodology presented in Appendix D to 40  CFR 132, a study of this type takes precedence over
other studies  in the development of a Tier I criterion.
                      Table 1-3. Summary of Chronic Avian Studies
Species
Mallard
Mallard
Mallard
Mallard
Kestrel
Pelican
LOAEL
(mg/kg/day)
1.16
2.91
0.58
1.45
0.39
0.028
NOAEL
(mg/kg/day)
0.116


0.58


Toxic Effect Observed
Eggshell thinning
Eggshell thinning
For DDE :
Reproductive effects
(Embryo mortality,
cracked eggs)
For DDT:
Reproductive effects
Eggshell thinning
Reproductive effects
Reference
Davison and Sell,
1975
Kolaja, 1977
Heath et al., 1 969
Heath et al., 1969
Peakall et al., 1973
Anderson et al.,
1975
///.  Avian Wildlife  Value Calculation
     The LOAEL is divided by a LOAEL to NOAEL uncertainty factor of 10, resulting in a
NOAEL for calculating avian wildlife values of 2.82 x 10"3 mg/kg/day based on reproductive
success.
     Most of the avian chronic laboratory studies (presented in Table 1-3) assessed effects of
DDT or metabolites on mallard ducks. Numerous accounts of DDT toxicity in birds observed in
the  field indicate piscivorous birds are among the most severely affected. This is further
supported by the acute toxicity data in which the bald eagle is the most sensitive of those tested.
It is unlikely, based on the toxicity database, that the pelican is the most sensitive species to the
lexicological impacts of DDT and its metabolites; therefore, a SSF of 0.1 is chosen as
appropriate.
                                                                                       1-7

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    The wildlife equation and input parameters are presented below.
                          NOAEL (avian)
                  BAF3 (Trophic Level 3)
                  BAF4 (Trophic Level 4)
                                    SSF
                                           2.82 //g/kg body weight/day
                                           1,000,000 //kg body weight
                                           3,000,000 //kg body weight
                                           0.1 (kingfisher, osprey and eagle)
    Values for body weights (WtJ, ingestion rate (FJ, and drinking rate (WJ for kingfisher, osprey
    and eagle are presented in Table D-2 of Appendix D to 40 CFR 132, and shown below.
                             WtA (kingfisher)
                               WtA (osprey)
                                 WtA (eagle)
                              FA (kingfisher)
                                 FA (osprey)
                                   FA (eagle)
                             WA (kingfisher)
                                WA (osprey)
                                 WA (eagle)
                                               0.15kg
                                               1.5kg
                                               4.5kg
                                               0.075 kg/day
                                               0.3 kg/day
                                               0.5 kg/day
                                               0.017 //day
                                               0.077 //day
                                               0.16  //day
                                                  + l<1-0)(FA(kinQWw)xBAF3)]

                                  	(2.82 i/g/kg/d x 0.110.15 kg	
                                  0.017 //d + [(1.01(0.075 kg/d x 1,000,000 //kg)]

                                                    0.56 pg//

                                             [NOAEL x SSF] x WtA(
-------
The geometric mean of these three avian wildlife values results in:
                    WV (avian)   =   glllnWVeungf^w) + InWVlMpray) + In WVlMgtell/3)

                    WV (avian)   =   e"'n-6po/' + mi.4pg// +

                    WV (avian)   =   0.87 pg/f.
iv. Sensitivity Analysis for Avian  Wildlife  Value
     The values of the various parameters used to derive the avian wildlife value presented above
represent the most reasonable assumptions. The purpose of this section is to illustrate the
significance of these assumptions and the variability in the avian wildlife value if other
assumptions are made for the values  of the various parameters from which the avian wildlife
value is derived. The intent of this section is to let the risk manager know, as much as possible,
the influence on the magnitude of the avian wildlife value of the assumptions made in its
derivation.
     Anderson et al. (1975) documented significant declines in DDT/DDE levels in the eggs and
prey of the brown pelicans, in addition to very moderate declines in the concentrations of PCBs,
mercury, and lead in their eggs (Anderson et al., 1977). The presence of these other pollutants in
the eggs  and prey may have contributed to the observed toxic effects on reproductive success
attributed to DDT. However, the levels of these contaminants remained constant over the
sampling period and were so low that this was deemed unlikely (Anderson et al., 1975; 1977).
Also, throughout the duration of the  study, declining DDT and  metabolite concentrations were
associated with increased eggshell thickness as well  as improved reproductive success.
     In estimating the hazard of DDT,  a SSF value of 0.1  was  used to account for possible
differences in sensitivity of pelicans compared to kingfisher, osprey, and eagles. This value was
based on the fact that these are all piscivorous species and piscivorous species appear more
sensitive to the lexicological effects of DDT and its metabolites. If an intermediate SSF of 0.3
(the geometric mean of 1 and 0.1) were used, with the LOAEL determined in the Anderson  et al.
(1975) study, the avian wildlife value would be of 2.6 pg/ instead of 0.87 pglt.
     The derivation of an avian wildlife value is based on the assumption that 100 percent of an
eagle's diet is composed of fish.  A study by Kozie and Anderson (1991) suggests that fish
comprise 97 percent of Lake Superior eagle diets, and mammals and birds each comprise 1.5
percent of eagle diets. Assuming the  metabolizable  energy in fish is approximately 1  kcal/g
(Palmer,  1988; Stalmaster and Gessaman,  1982) and the typical eagle consumes about 500 g of
fish per day (Technical Support Document for Wildlife Criteria, Appendix to the Preamble to 40
CFR 132), an eagle has a daily energy requirement of 500 kcal/day. The energy content for  birds
is 2 kcal/g (a value derived for mallards; Stal master and Gessaman, 1982). Applying the
conservative assumptions that the bioaccumulation in mammals  would be equivalent to that in
Trophic Level 4 fish and the  caloric value would be the same for mammals and fish, an eagle diet
consisting of 1.5 percent fish-eating birds and 98.5  percent fish would result in a daily intake of
approximately 7.4  g of bird and 480  g of fish to meet the daily energy requirement of 500
kcal/day. Braune and Nordstrom (1989) have reported that DDE bioaccummulates in Lake
Ontario herring gulls at a level approximately 85 times higher than that observed in alewife.
Therefore, dietary  exposure of eagles to DDE would be higher if piscivorous birds comprise a
portion of their diet. The DDE exposure to eagles eating 7.4 g of piscivorous birds a day would
be approximately 2.4 times higher than an exposure associated with a 100 percent fish diet. Such
an analysis would result in a bald eagle wildlife value of 0.38 pg/l, and an avian wildlife value of
0.67 pglt compared to 0.87
                                                                                        1-9

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IV.  Great Lakes  Wildlife Criterion
          The Tier I Great Lakes Wildlife Criterion for p,p -Dichlorodiphenyltrichloroethane (DDT)
    and metabolites is determined by the lower of the mammalian wildlife value (270 pg/l) and the
    avian wildlife value (0.87 pg/l). The avian wildlife value was determined to be approximately 4
    orders of magnitude smaller that the mammalian wildlife value. Therefore, the Great Lake
    Wildlife Criterion for DDT and metabolites is 0.87 pg/l.


    /. Discussion  of Uncertainties
          Wildlife populations inhabiting the Great Lakes Basin would not be impacted from the intake
    of drinking water or prey taken from surface water containing total DDT in concentrations of
    0.87 pg/l, based on available exposure, toxicity and bioaccumulation information, and uncertainty
    factors applied  to account for data gaps and the variability inherent in the DDT risk assessment.
    Criteria for other ecoregions may require an analysis of different wildlife species with different
    diets and body  masses. In addition,  the bioaccumulation factors in this analysis were based on an
    analysis specific for the Great Lakes; different bioaccumulation factors may be more appropriate
    for other waterbodies.
          Generic assumptions were made in assessing the hazards of DDT and its metabolites to
    wildlife populations through the use of LOAELs and NOAELs for reproduction and development.
    The use of these levels assumes no hazards to wildlife populations would result from the direct
    exposure of individuals to DDT and its metabolites. However, it could be argued that some
    increase in density independent mortality, or decrease in density independent reproductive
    success, which  could be attributable to exposure to DDT or its metabolites, could be incurred
    without impacting the population dynamics of a species. In general, well-validated population
    models do not yet exist for the species analyzed,  and it is difficult to estimate the extent of
    mortality or reproductive failure that could be incurred. In addition, the interaction of additional
    chemical as well as non-chemical  stressors on wildlife population responses is also poorly
    resolved at this time.


V.  References
AIsop, F. J.  1972. Eggshell thickness from red-winged blackbird (Ageloius phoeruceus) populations with different
    exposures to DDT. Dissertation Abstracts  33 (11):5571-B.
Anderson, D. W., J. R. Jehl, R. W. Risebrough, L. A. Woods, L. R. Dewecse and W.G. Edgecombe. 1975. Brown
    pelicans: Improved reproduction off the southern California coast. Science 190:806-808.
Anderson, D. W., R.  M. Jurek and J.  O. Keith. 1977. The status of brown pelicans at Anacapa Island in 1975. Calif.
    Fish and Game  1:4-10.
Bernard, R.F. 1963. Studies on the effects of DDT on birds. Biological Series Mi. State U. Museum 2(3): 159-191.
Bloom. 1973. Seasonal variation in body weight of sparrow hawks in California. Western Bird Bander 48:17-19.
Braune, B. M. and  R. J. Norstrom.  1989. Dynamics of organochlorine compounds in herring gulls:in. Tissue
    distribution and bioaccumulation in Lake Ontario gulls. Environ.  Toxicol. Chem. 8:957-968.
Davison, K.  L. and J. L.  Sell. 1973. DDT and dieldrin effects on mallard ducks. Federal Proceedings 32(3):320.
Davison, K.  L. and J. L.  Sell. 1974. DDT thins shells of eggs from mallard ducks maintained on ad libitum or
    controlled-feeding regimens. Arch. Environ. Con tarn. Toxicol. 2:222-232.
Delnicki, D.  and K. J. Reinecke.  1986. Mid-winter food uses and body weights of mallards and wood ducks in
    Mississippi. J. Wildl. Manage. 50:43-51.
Dunning, J. B. 1984.  Body Weights of 686 North American Birds. Monograph #1. Western Bird Banding Association.
Fltzhugh, O. 1948.  Use of DDT insecticides on food products. Industrial and Engineering Chemistry 40(4):704-705.
Gilbert, F. 1969. Physiological effects  of natural DDT residues and metabolites on ranch mink.  J. Wildlife Manage.
    33(4):933-943.
Haegele M.A., and  R.H. Hudson. 1974. Eggshell thinning and residues in mallards one year after DDE exposure.
    Arch, of Environ. Contain, and Toxicol. 2(4):356-363.
Heath, R. G., J. W. Spann and J. F. Kreitzer. 1969. Marked DDE impairment of mallard reproduction in controlled
    studies. Nature  224:47-48.
Kolaja, G. J. 1977. The effect on DDT, DDE and their sulfonated derivatives on eggshell formation in the mallard
    duck.  Bull. Environ. Con tarn. Toxicol.  17(6): 697-701.
Kolaja, G. J., and D.  E. Hinton. 1977. Effects of DDT on eggshell quality and calcium adenosine triphosphatase. J.
    Toxicol. Environ.  Health 3:699-704.

_                                                                     _

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Kozie K.D. and R.K. Anderson. 1991. Productivity, diet, and environmental contaminants in bald eagles nesting near
     the Wisconsin shoreline of Lake Superior. Arch. Environ. Contain. Toxicol. 20:41-48.
Luig, E.P., A. Nelson, G. Fitzhugh, and F. Kunze. 1950.  Liver cell alteration and DDT storage in fat of the rat
     induced by dietary levels of 1 to 50 ppm DDT. Pharmacol. Exp. Therap. 98:268-273.
Longcore, J. R., and F. B. Samson. 1973. Eggshell breakage by incubating black ducks fed DDE. I. Wildl. Manage.
     37(3): 390-394.
Mitjavila, S., G. Carrera, R.-A. Boigegrain and R. Derache. 1981.  I. Evaluation of the toxic risk of DDT in the rat:
     during accumulation. Arch. Environ. Contain. Toxicol. 10:459-469.
National Institute for Occupational Safety and Health (NIOSH).  1992. General toxicity file for DDT (CAS No. 59-
     29-3) in Registry of Toxic Effects of Chemical Substances (RTECS datahise, available only on microfiche or as an
     electronic database). Cincinnati, OH.
Palmer, R.S. Editor. 1988. Handbook of North American Birds, Vol. 4. Yale University Press. 433 pp.
Peakall, D. B., I. L. Lincer, R. W. Riaebrough, I. G. Pritchard and W. B. Kinter. 1973.  DDE-induced egg-shell
     thinning: Structural and physiological effects in three species. Comp. Gen. Pharmacol. 4:305-313.
Porter, R.D. and S.N. Wiemeyer.  1972. DDE at low dietary levels  lolls captive American kestrels. Bull.  Environ.
     Contain. Toxicol. 8(4): 193-199.
Risebrough, R. W. and D. W. Anderson.  1975. Some effects of DDE and PCB on mallards and their eggs. I. Wildl.
     Manage. 39(3):508-513.
Stalmaster, M. V., and I. A. Gessaman.  1982. Food consumption and energy requirements of captive bald eagles. I.
     Wildl. Manage.  46:646-654.
Stickel, W.H., L.F.  Stickel, R.S. Dyriand and D.L. Hughes. 1984.  DDE in birds: Lethal residues  and loss rates.
     Arch. Environ. Contain. Toxicol.  13:1-6.
Stickel, L.F., N.I. Chura, P.A. Stewart, C.M. Menzie, R.M. Prouty and W.L. Reichel. 1966. Bald eagle pesticide
     relations. Transactions of the 31st North American Wildlife and  Natural Resources Conference 31:190-201.
U.S. Environmental Protection Agency. 1980, Nov. 28. Water Quality Criteria Documents, Availability. Federal
     Register 45:79318-79378.
Van Vdtzen, W. and P. Kreitzer. 1975. The toxicity of p,p'-DDT to the clapper rail. J. Wildl. Manage.  39(2):305-
     309.
                                                                                                         1-11

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CHAPTER 2
Tier  I  Wildlife Criteria  for  Mercury
(Including  Methylmercury)
                               Contents


I.   Literature Review	   2-1

H.  Calculation of Mammalian Wildlife Value	   2-1

    i.  Acute Toxicity Studies	   2-1

    ii.  Chronic Toxicity Studies	   2-1

    iii.  Mammalian Wildlife Value Calculation	   2-2

    iv.  Sensitivity Analysis for Mammalian Wildlife Value  	   2-3

ffl.  Calculation of Avian Wildlife Value 	   2-4

    i.  Acute Toxicity Studies	   2-4

    ii.  Chronic Toxicity Studies	   2-5

    iii.  Avian Wildlife Value Calculation 	   2-7

    iv  Sensitivity Analysis for Avian Wildlife Value	   2-8

IV.  Great Lakes Wildlife Criterion	   2-9

    i.  Discussion of Uncertainties  	   2-9

V.  References 	   2-9

-------
 Tier  I  Wildlife  Criteria  for  Mercury
 (Including   Methylmercury)
 I. Literature  Review
         A review of the available literature on the environmental cycling, fate, and toxicity of
    mercury and mercury compounds indicates that criterion derivation for mercury is most
    appropriately based on methylmercury. A review of mammalian and avian toxicity data for
    methylmercury was based on literature identified through computer-based (CAS and BIOSES), as
    well as manual, searches. A total of 27 references were screened; those references which were
    reviewed in detail are cited in Section V and primarily include those which contained dose-
    response data.


 II. Calculation of Mammalian Wildlife Value
    /. Acute Toxicity Studies
         Methylmercury and other organomercury compounds are the most toxic forms of mercury to
    mammals. Methylmercury affects the central nervous system, resulting in sensory, visual, and
    auditory impairment. Experimentally induced acute mercury poisoning in mule deer (Odocoileus
    hemionus) was characterized by belching, bloody diarrhea, piloerection (i.e., the hair was more
    erect than usual), and loss of appetite. The kidney appears to be the critical organ affected in
    adult mammals as a result of the rapid degradation of phenylmercurials and
    methoxyethylmercurials to inorganic mercury compounds. In the fetus, the brain is the principal
    target (Eisler,  1987).
        The differential toxicity of the different forms of mercury is exemplified by the results of a
    study by Aulerich et al. (1974). Using adult mink (Mustela visori), dietary exposure to 5 ppm of
    methylmercury was found to be lethal in about 1 month, while exposure to 10 ppm of mercuric
    chloride did not produce adverse effects over 5 months.
        Death in  sensitive mammalian species has been associated with daily organomercury doses
    of 0.1 to 0.5 mg/kg body weight and 1 to 5 mg/kg in the diet. Larger mammals such as mule
    deer and harp seals (Pagophilus groenlandicus) appear  to be more resistant to the toxic effects of
    mercury than smaller mammals. Mule deer had organomercury LDjo values of 17.88 mg/kg body
    weight, and  all harp seals exposed to mercury at 25 mg/kg body weight died  within 28 days  of
    dietary exposure. Doses of 1.0 mg/kg in the diet produced death in all experimental mink within
    2 months of exposure and a dose  > 2.0 mg/kg killed all experimental river otters (Lutra
    canadensis); in cats (Felis domesticus), convulsions and reductions in survival were associated
    with organomercury exposure at 0.25 mg/kg body weight for 90 days (Eisler, 1987).

    //'. Chronic  Toxicity Studies
        Wobeser  et al. (1976a) examined the effects of organic and inorganic mercury on mink.
    Wobeser et al. (1976) fed adult female and juvenile ranch mink rations consisting of 50 and 75
    percent fish that contained 0.44 ppm total mercury over a 145-day period. The corresponding
    concentrations  of inorganic mercury in the diet are 0.22 and 0.33 ppm. No clinical or
    pathological  signs of intoxication were observed at these exposure concentrations, suggesting a

_                                                                                    _

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    NOAEL of 0.33 ppm. Using the mink body weight of 1.0 kg and food ingestion rate of 0.15
    kg/day provided in the methodology (Appendix D to 40 CFR 132), the NOAEL from this study
    is 0.05 mg/kg/day.
         In a subsequent dose-response study, Wobeser et al. (1976a) fed adult female mink rations
    treated with methylmercury chloride at concentrations of 1.1, 1.8, 4.8, 8.3, and 15.0 ppm total
    mercury for 93 days. Mink exposed to dietary mercury concentrations of 1.8 ppm and greater
    developed signs of clinical intoxication (anorexia, ataxia, and death). The time to onset of the
    toxic effects was directly related to the mercury content of the ration, and therefore, to the total
    dose. Pathological alterations in the nervous system were observed at the 1.1 ppm concentration,
    although they were not associated with any obvious clinical evidence of toxicity. Because these
    lesions were observed hi the nervous systems of animals receiving 1.1 ppm mercury, the authors
    argued that distinct clinical signs of toxicity would have developed in animals at that dose had the
    experimental period been longer. Based on these results, the NOAEL for mortality in  mink fed
    methylmercury is 1.1 ppm, and the LOAEL is 1.8 ppm. Using the mink body weight  and food
    ingestion rate presented above, the LOAEL is 0.27 mg/kg/day, and the NOAEL is 0.16
    mg/kg/day.
        The NOAEL of 0.16 mg/kg/day elemental mercury (methylmercury chloride) reported by
    Wobeser et al. (1976a) is used to calculate a mammalian-based mercury wildlife value (WV).
    This study consists of repeated oral exposures for over a 90-day period using a mammalian
    wildlife species, and therefore meets the criteria for an appropriate study for wildlife criteria
    development as described in Appendix D to 40 CFR 132.

    Hi. Mammalian  Wildlife Value Calculation
        A subchronic  to chronic conversion factor of 0.1  is used because of the relatively short
    duration of the study and the time course of histopathological changes observed in mink fed 0.16
    mg/kg/day methylmercury (Wobeser et al. 1976a). This results in an adjusted NOAEL of 0.016
    mg/kg/day.
        In  calculating a Tier I wildlife value, a species sensitivity factor (SSF) within the range of 1
    to 0.01 is recommended  in Appendix D to 40 CFR 132 to accommodate differences in
    lexicological sensitivity between the experimental animal and the mink and river otter. A SSF of
    1 is used because the NOAEL is based on a study using mink as  the test species.
        The bioaccumulation factors (BAFs) relate concentration of methylmercury in fish tissue to
    the concentration of total mercury in the water column. The methylmercury BAFs for Trophic
    Level 3 and Trophic Level 4 are derived based on the procedure specified in the Great Lakes
    Water Quality Initiative guidance on bioaccumulation, found in Appendix B to 40 CFR 132,
    entitled Methodology for Development of Bioaccumulation Factors.

    Input parameters for the  wildlife equation are presented below.
                      NOAEL (mammalian)   =   0.016 mg/kg/day

                     BAF3(Trophic Level 3)   =   60,000 I /kg body weight

                     BAF4(Trophic Level 4)   =   130,000 t/kg body weight

                                     SSF   =   1 (mink and otter)
2-2

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 WtA, FA, and WA for mink and river otter are presented in Table D-2 of the methodology
 presented in Appendix D to 40 CFR 132 and are shown below.
WtA(mink)
WtA(otter)
FA(mink)
FA(otter)
WA(mink)
WA(otter)
= 1 .0 kg
= 8.0 kg
= 0.15kg/d
0.9 kg/d
0.099 l/d
= 0.64 t/d
The wildlife equations and calculations of mammalian wildlife values are summarized below.
 WV (mink)


 WV (mink)


 WV (mink)


 WV (otter)



 WV (otter)


 WV (otter)
                        [NOAEL x SSF1 x WtA(minl[,
                     WA(mink) + N1.0)(FA(mH[)xBAF3)]


              _ (0.016 mg/kg/d x 1) 1.0kg _
               0.099 l/d +  [(1.01(0.15 kg/d x 60,000 I/kg)]
                              1,800pg/l
[NOAEL x SSF] x WtAloror| _
                           x BAF4)]
           WA(omrl + H0.5)(FA(omf) x BAF3) + (

                       (0.01 6 mg/kg/d x 1) 8.0kg
0.64 l/d +  [{0.5M0.90 kg/d x 60,000 I/kg)  + (0.51(0.90 kg/d x 130,000 I/kg)]
                                                1 ,500 pg/l
The geometric mean of these two mammalian wildlife values results in:
                          WV (mammalian)   =   e111"^"1*1 + lnWVIOtM"l/2)

                          WV (mammalian)   =   e"ln 1-800 P8" + ln 1-600 p"m)

                          WV (mammalian)   =   1,600pg/.
iv. Sensitivity Analysis for Mammalian Wildlife Value
     The values of the various parameters used to derive the mammalian wildlife value presented
above represent the most reasonable assumptions. The purpose of this section is to illustrate the
significance of these assumptions and the variability in the mammalian wildlife value if other
assumptions are made for the values of the various parameters from which the mammalian
wildlife value is derived. The intent of this section is to let the risk manager know, as much as
possible, the influence on the magnitude of the mammalian wildlife value of the assumptions
made hi its derivation.
     In deriving the mammalian wildlife value for mercury, it was assumed that 100 percent of
the mink diet was comprised of fish, although this may not necessarily be the case. This
assumption may lead to an overestimate of mercury exposure for those organisms that are not
                                                                                       2-3

-------
    primarily feeding on fish and aquatic invertebrates. However, as indicated in the Technical
    Support Document for Wildlife Criteria (Appendix to the Preamble to 40 CFR 132), the fish
    content of a mink diet can vary from less than 50 percent to the 100 percent assumed in the mink
    wildlife value derivation presented above. If it were assumed only 50 percent of a mink's diet was
    from aquatic resources and the remaining 50 percent of the diet was uncontaminated, the
    estimated mercury exposure would be reduced by  a factor of 2. The mink wildlife value would be
    3600 pg/l, and the mammalian wildlife value would be 2,600 pg/l, rather than the mammalian
    wildlife value of 1,600 pg/l.
III.  Calculation of Avian Wildlife Value
    /. Acute  Tox/c/ty Studies
         Methylmercury has been shown to be more toxic to avian species than inorganic mercury.
    Acute oral toxicity of methylmercury produced LD,, values ranging from 2.2 to 23.5 mg/kg for
    mallards (Anasplatyrhynchos), 11.0 to 27.0 mg/kg for quail (Coturnix), 14.4 to 33.7 for
    Japanese quail (Coturnix japonica), and 23.8 mg/kg for northern bobwhite (Colinus virginianus).
    Inorganic mercury produced LDM values of 26.0 to 54.0 mg/kg in quail, and 31.1 in Japanese
    quail (Eisler, 1987). The LDW values for avian species are summarized in Table 2-1.
    Furthermore, some birds poisoned by inorganic mercury recovered after treatment was
    withdrawn, while chicks mat were fed methylmercury usually died, even after the treated feed
    was removed.
         Mercury poisoning in birds is characterized by muscular incoordination, falling, slowness,
    fluffed feathers, calmness, withdrawal, hyperactivity, hypoactivity, and eyelid drooping (Eisler,
    1987). Following acute oral exposures, signs of mercury poisoning have been observed within 20
    minutes after administration in mallards, to 2.5 hours after administration in pheasants. Death
    occurred between 4 and 48 hours in mallards and 2 and 6 days in pheasants (Hudson et al.,
    1984).
                    Table 2-1. Summary of Avian Acute Oral Toxicity Values
Mercury Form
Inorganic
Organic
Species
Japanese quail (Coturnix japonica)
Coturnix (Cotrunix coturnix)
Chukar (Alectoris chukar)
Mallard (Anas platyrhynchos}
Northern bobwhite (Colinus virginianus)
Coturnix (Cotrunix coturnix)
Japanese quail (Coturnix japonica)
Rock dove (Columba I/via)
Fulvous whistling duck (Dendrocygna bicolor)
Domestic chicken (Callus domesticus)
House sparrow (Passer domesticus)
Gray partridge (Perdix)
Ring-necked pheasant (Phasianus colchicus)
Prairie chicken (Tympanucus cup/do)
LD60 (mg/kg)
14.1 -33.7
2956 - 5086
26.9
2.2-75.7
23.8
4-27
14.4-33.7
22.8
37.8
60
12.6-37.8
17.6
11.5 - 112
11.5
 Sourc*: Eisler (1987).
2-4

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//.  Chronic Tox/c/ty Studies
     Fimreite (1970) raised two-week old leghorn cockerel chicks (Callus) on commercial feed
containing methylmercury dicyandiamide at concentrations of 0, 6, 12, and 18 ppm for 21 days.
A significant increase in mortality was observed at the highest concentration of methylmercury
(18 ppm); however, mortality in chicks maintained at 6 or 12 ppm was not significantly different
man mat in the control group. Hence, the NOAEL for mortality is 12 ppm. Growth was
significantly reduced in chicks maintained on mercury treated feed, suggesting a LOAEL for
growth of 6.0 ppm. Using a juvenile chicken body weight of 0.8 kg and a food consumption rate
of 0.14 kg/day (NIOSH,  1991) the NOAEL for mortality resulting from ingestion of
methylmercury in chicken can be calculated to be 2.1 mg/kg/day,  and the LOAEL for growth can
be  calculated to be 1.1 mg/kg/day.
     In another study, Fimreite (1971) exposed ring-necked pheasants to grain treated with a seed
dressing containing 2.5 percent methylmercury dicyandiamide at doses of mercury  equivalent to
approximately 0.69 mg/kg/day, 0.37 mg/kg/day, and 0.18 mg/kg/day for 12 weeks. The laying
hens showed no acute symptoms of mercury poisoning; however, adverse effects on reproduction
of the pheasants were observed at all dose levels. Reduced hatchability was the most significant
effect, while reduced egg production and increased numbers of shell-less eggs were also
observed. Among the eggs that hatched, chick mortality appeared to be only slightly increased.
The results of this study suggest a LOAEL for total mercury as methylmercury effects on
reproduction in pheasants of 0.18 mg/kg/day.
     Scott (1977) provided white leghorn laying hens with methylmercury chloride at dietary
concentrations of 0, 10, and 20 ppm, and inorganic mercury (mercuric sulfate)  at concentrations
of 100 and 200 ppm. Methylmercury at 10 and 20 ppm was found to severely impact egg
production and weight, fertility of eggs, hatchability of fertile eggs, and eggshell strength. Dietary
levels of 100 or 200 ppm inorganic mercury had little affect on egg production, hatchability, shell
quality, morbidity, and mortality. The LOAEL for reproductive effects of methylmercury in white
leghorn chickens obtained from this study is 10 ppm. Using a chicken body weight of 1.66 kg
(Lillie et al., 1975; and personal communication with Dr. Wayne Kunzel, Poultry Science
Department, University of Maryland) and a food ingestion rate of 0.81 kg/day derived from the
allometric relationship presented  in the methodology (Appendix D  to 40 CFR 132), the LOAEL
for reproductive effects of methylmercury is 4.9  mg/kg/day.
     In a series of studies Heinz (1974, 1975, 1976, 1976a, 1979) assessed the effects of dietary
methylmercury over three generations of mallard ducks. Adult mallards and ducklings were
maintained on a commercial feed treated with methylmercury dicyandiamide at concentrations
equivalent to 0, 0.5, and 3.0 ppm elemental mercury (the nominal  treatment levels  were
confirmed by atomic absorption analysis for elemental mercury).
     Initially, adult mallard ducks were maintained for a period of up to 21 weeks  on the treated
diet (Heinz,  1974). There were no consistently large differences in eggshell thickness among the
three groups. Egg production stopped earlier among the 3 ppm group than among the 0.5 ppm or
control group. Hatching success and hatchling viability, as measured by the number of normal
hatchlings and survival of hatchlings through one week, were significantly reduced  in the 3.0 ppm
group but not in the 0.5 ppm group. These results indicate a LOAEL for reproduction of 3.0 ppm
and a NOAEL of 0.5 ppm.
     Reproduction in first and second generation ducks was also evaluated (Heinz,  1976, and
1976a). In the first generation, no significant reproductive effects were reported based on an
assessment of percent cracked eggs, egg production, or the number of eggs producing normal
hatchlings. However, offspring survival to  1-week was significantly lower in the 3.0 ppm
treatment group, but not in the 0.5  ppm group. In the second generation of  parents  fed dietary
mercury, abnormal egg-laying behavior, impaired reproduction, and slowed growth of ducklings
were observed in the ducks fed 0.5 ppm mercury (Heinz,  1976a). These results suggest a LOAEL
of 3.0 ppm and a NOAEL of 0.5 ppm for offspring survival in the first generation, and a
LOAEL of 0.5 ppm for reproductive effects in the second generation of offspring.
                                                                                       2-5

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         The reproductive and behavioral effects of methylmercury during the third breeding season
    are reported in Heinz (1979). During the final year of the study, ducks were only provided with 0
    or 0.5 ppm dietary elemental mercury in feed. The results of this study, combined with the earlier
    investigations (Heinz, 1974, 1975, and  1976) provide dose-response relationships over three
    generations of mallard ducks. Heinz (1979) found that third generation bens fed methylmercury at
    0.5 ppm laid fewer sound eggs man controls. Fewer sound eggs were also observed in the 0.5
    ppm group when data were combined across all  generations. The percent of incubated eggs
    producing normal hatchlings and the percent of normal hatchlings surviving 1 week were not
    significantly reduced by dietary methylmercury exposures. Only during the second generation was
    the number of 1-week old ducklings produced significantly reduced; however, when pooling these
    data with the results of the first and  third generations, a significant effect was detected.
         Heinz (1975, 1976, 1976a, and 1979) examined approach and avoidance in mallard
    ducklings maintained on treated diets and on the hatchlings born in the second and third breeding
    season.  The behavior tests were designed to measure the approach response to maternal calls, and
    the avoidance response to a frightening stimulus. Among the initial group of ducklings, alteration
    of the approach and avoidance behavior was observed at the 0.5 ppm level. During the  second
    generation, the ducklings of parents  who were fed 3 ppm mercury were hyper-responsive
    compared to controls and ducklings  from parents fed 0.5 ppm. Altered duckling approach
    responses were observed in the third generation  and when data were pooled over all generations
    at die low treatment level. Avoidance behavior of ducklings was not significantly altered within
    any generation. However, when data were pooled across generations, a significant effect was
    obtained at the low dietary concentration of 0.5 ppm. During the second generation and for the
    combined data across generations, Heinz (1979) reported that hens laid a significantly higher
    percentage of eggs outside the nestbox.
         Based on the observed adverse reproductive and behavioral effects  across the three
    generations, a  LOAEL of 0.5 ppm elemental mercury, as methylmercury, can be inferred. In the
    multi-generational mallard study, a food consumption rate for mallard ducks was reported to be
    128 g/kg/day based on the combined data for controls from the second and third generations.
    Multiplying the 0.5 ppm dietary mercury LOAEL by the food consumption rate results  in a
    LOAEL of 0.064 mg/kg/d.
    The results of the studies described previously are summarized in Table 2-2.

                          Table 2-2. Summary of Avian Chronic Studies
Species
Chicken
(juvenile)
Pheasant
Chicken
Mallard
LOAEL
(mg/kg/day)
1.1

0.18
4.9
0.064
NOAEL
(mg/kg/day)

2.1



Toxic Effect Observed
Growth
Mortality
Reproduction
Reproduction
Reproduction,
behavior
Reference
Fimreite, 1970
Fimreite, 1971
Scott, 1977
Heinz, 1974,
1975, 1976,
1976a, 1979
         The results of the Heinz (1974, 1975, 1976,  1976a, and 1979) multigeneration studies of the
    effects of methylmercury on mallard ducks were judged to be the most appropriate for derivation
    of the avian wildlife value. These studies provide a chemical-specific dose-response curve over
    three generations with explicitly quantified effects  on reproduction and behavior. These effects
    clearly have potential consequences on populations of mallards exposed to methylmercury.
2-6

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Hi. Avian Wildlife Value Calculation
     The LOAEL was divided by a LOAEL to NOAEL uncertainty factor of 2 because the
LOAEL appeared to be very near the threshold for dietary effects. Applying this factor to the
LOAEL presented previously gives a NOAEL for calculating avian wildlife values of 3.2 x 10*2
mg/kg/day.
     Given the limited number of species for which dose-response data is available on the chronic
effects of mercury and the lack of avian NOAEL data in these studies, a SSF of 0.1 is used to
calculate a wildlife value for kingfisher, osprey, and eagle.
     The wildlife equation and input parameters are presented below. The BAFs relate
concentration of methylmercury in fish tissue to the concentration of total mercury in the water
column.
                   NOAEL (avian)
            BAF3 (Trophic Level 3)
            BAF4 (Trophic Level 4)
                             SSF
32 //g/kg body weight/day
60,000 I/kg body weight
130,000 //kg body weight
0.1  (kingfisher, osprey, and eagle)
Values for body weights (WtJ, food ingestion rate (FJ, and water ingestion rate (WJ for
kingfisher, osprey, and eagle are presented in Table D-2 of the methodology document (Appendix
D to 40 CFR 132) and shown below.
                       WtA(kingfisher)
                         WtA( osprey)
                           WtA(eagle)
                        FA(kingfisher)
                           FA(osprey)
                            FA(eagle)
                       WA(kingfisher)
                          WA(osprey)
                           WA(eagle)
    0.15 kg
    1.5kg
    4.5kg
    0.075 kg/d
    0.3 kg/d
    0.5 kg/d
    0.017 f/d
    0.077 /d
    0.16 f/d
                                                                                     2-7

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    Calculations of avian wildlife values are summarized below.
                 WV (kingfisher)



                 WV (kingfisher)


                 WV (kingfisher)



                   WV (osprey)



                   WV (osprey)


                   WV (osprey)

                     WV (eagle)



                     WV (eagle)



                     WV (eagle)
    (NOAEL x SSF) x WtA(hingfirt,,WV (eagle)
                Id.OHF^,^, x BAF3)l
      (32//g/kg/d xO.1) 0.15 kg
0.017 lid + [(1.0H0.075 kg/d x 60,000 //kg)]

                 100pg//


       (NOAEL x SSF) x WtA(
-------
    conservative assumptions that the bioaccumulation in mammals would be equivalent to that in
    Trophic Level 4 fish and the caloric value would be the same for mammals and fish, an eagle diet
    consisting of 1.5 percent fish-eating birds and 98.5 percent fish would result in a daily intake of
    approximately 7.4 g of bird and 480 g of fish to meet the daily energy requirement of 500
    kcal/day.  If methylmercury biomagnifies in fish-eating birds as has been observed for 2,3,7,8-
    TCDD, DDT, and PCBs (see Braune and Norstrom, 1989), dietary exposure of methylmercury to
    eagles would be higher if piscivorous birds comprise a portion of their diet than if the diet were
    composed of 100 percent fish. However, a quantitative estimate of an avian wildlife value
    adjusted for this additional exposure <"\n not be determined because the  empirical data on
    bioaccumulation of mercury at higher trophic levels is not available.


IV.  Great Lakes Wildlife  Criterion
         The  Tier I Great Lakes Wildlife Criterion for mercury is determined by the lower of the
    mammalian wildlife value (1,600 pg/l) and the avian wildlife value (180 pg/l)- The Wian wildlife
    value is one order of magnitude lower than the mammalian value. Therefore the Great Lakes
    Wildlife Criterion for mercury is 180 pg/f.


    /. Discussion of Uncertainties
         Wildlife populations inhabiting the Great Lakes Basin would not be impacted from the intake
    of drinking water or prey taken from surface water containing total  mercury in concentrations of
    180 pg/l, based on available exposure, toxicity and bioaccumulation information, and uncertainty
    factors applied to account for data gaps  and the variability inherent in the mercury risk
    assessment. Criteria for other ecoregions may require an analysis of different wildlife species with
    different diets and body masses than were used for the Great Lakes  Basin. In addition, the
    bioaccumulation factors in this analysis were based on an analysis specific for the Great Lakes;
    different bioaccumulation factors may be more appropriate for other waterbodies.
         Finally, generic assumptions were  made in assessing the hazards of mercury to wildlife
    populations through the use of LOAELs and NOAELs for reproduction and development. The use
    of these levels assumes no hazards to wildlife populations would result from the direct exposure
    of individuals to mercury. However, it could be argued that some increase in density independent
    mortality, or decrease  in density  independent reproductive success, which could be attributable to
    mercury exposure, could be incurred without impacting the population dynamics of a species. In
    general, well-validated  population models do not yet exist for the species analyzed, and it is
    difficult to estimate the extent of mortality or reproductive failure that could be incurred. In
    addition, the interaction of additional chemical as well as non-chemical stressors on wildlife
    population responses is also poorly resolved at this time.


V. References
Aulerich, RJ.,  R.K. Ringer and S. Iwamoto. 1974. Effects of dietary mercury on mink. Arch. Environ. Co item.
    Toxicol. 2(1):43-51.
Braune, B. M.  and R.  I. Norstrom.  1989.  Dynamics of organochlorine compounds in herring gulls: Id. Tissue
    distribution and bioaccumulation in Lake Ontario gulls. Environ. Toxicol. Chem. 8:957-968.
Eisler, R. 1987. Mercury Hazards to Fish,  Wildlife, and Invertebrates: A Synoptic Review. U.S. Fish Wildl. Serv.
    Biol. Rep. 85(1.10). 90pp.
Rmrdte, N. 1970.  Effects of methyl mercury treated feed on the mortality and growth of leghorn cockerels. Can. J.
    Anim. Sci.  50:387-389.
Fimreite, N. 1971.  Effects of methylmercury on ring-necked pheasants. Canadian Wildlife Service Occasional Paper
    Number 9. Department of the Environment. 39 pp.
Heinz, G.H. 1974.  Effects of low dietary levels of methyl mercury on mallard .reproduction. Bull. Environ. Contain.
    Toxicol. 11:386-392.
Heinz, G.H. 1975.  Effects of methylmercury on approach and avoidance behavior of mallard ducklings. Bull.  Environ.
    Contain. Toxicol. 13:554-564.
Heinz, G.H. 1976.  Methylmercury: Second-year feeding effects on mallard reproduction  and duckling behavior. J.
    Wildl. Manage. 40(1):82-90.

                                                                                               

-------
Heinz, G.H. 1976*. Methylmercury: Second-generation reproductive and behavioral effects on mallard ducks. J. Wildl.
     Manage. 40(4):710-715.
Heinz, G.H. 1979. Methylmercury: Reproductive and behavioral effects on three generations of mallard ducks. I.
     Wildl. Manage. 43:394-401.
Hudson, R.H., R.K. Tucker and M.A. Haegele. 1984.  Handbook of Toxicity of Pesticides to Wildlife. U.S. Fish
     Wildl. Serv.  Resour. Publ. 153. 90 pp.
Kozie, K.D. and  R.O. Anderson.  1991. Productivity, diet, and environmental contaminants in bald eagles nesting near
     the Wisconsin shoreline of Lake Superior. Arch. Environ. Con tarn. Toxicol. 20:41-48.
Lillie, R. J., H. C. Cecil, J. Bitman and G. F. Fries. 1975.  Toxicity of certain polychlorinated  and polybrominated
     biphenyls on reproductive efficiency of caged chickens.  Poultry Sci. 54:1500-1555.
National Institute for Occupational Safety and Health (NIOSH).  1991. Registry of Toxic Effects of Chemical
     Substances (RTECS database, available only on microfiche or as an electronic database). Cincinnati, OH.
Palmer, R.S. Editor. 1988. Handbook of North American Birds. Vol. 4.  Yale University Press. 433 pp.
Scott, M. L. 1977. Effects of PCBs, DDT, and mercury compounds in chickens and Japanese quail. Federation
     Proceedings.  36:1888-1893.
Stalmaster, M. V., and J. A. Gessaman. 1982. Food consumption and energy requirements of captive bald eagles. J.
     Wildl. Manage. 46:646-654.
Wobeser, G., N.D. Nielsen and B. Schiefer.  1976. Mercury and Mink I: The use of mercury contaminated fish as a
     food for ranch mink. Can. J. Comp. Med. 40:30-33.
Wobeser, G., N.D. Nielsen, and B. Schiefer. 1976a. Mercury and Mink II: Experimental methyl mercury intoxication.
     Can. J. Comp. Med. 40:34-45.
2-10

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CHAPTER 3
Tier I Wildlife  Criteria  for 2,3,7,8-
Tetrachlorodibenzo-p-dioxin  (2,3,7,8-TCDD)
                             Contents


I.  Literature Review	  3-1

I.  Calculation of Mammalian Wildlife Value	  3-1

   i.  Acute Toxicity	  3-1

   ii. Chronic Toxicity	  3-2

   iii. Mammalian Wildlife Value Calculation	  3-4

   iv. Sensitivity Analysis for Mammalian Wildlife Value 	  3-5

ID. Calculation of Avian Wildlife Value 	  3-6

   i.  Acute Toxicity	  3-6

   ii. Chronic Toxicity	  3-6

   iii. Avian Wildlife Value Calculation 	  3-7

   iv. Sensitivity Analysis of Avian Wildlife Value  	  3-8

IV. Great Lakes Wildlife Criterion	  3-9

   i.  Discussion of Uncertainties 	  3-9

V.  References 	  3-10

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Tier  I Wildlife  Criteria  for  2,3,7,8-
Tetrachlorodibenzo-p-dioxin
(2,3,7,8-TCDD)
I.  Literature Review
       A review of mammalian and avian toxicity data for 2,3,7,8-TCDD was based on literature
   received through computer-based (CAS and BIOSES) as well as manual searches. A total of 13
   references were screened; those references that were reviewed in detail are cited in Section V,
   and primarily include those that contain dose-response data.


II. Calculation of Mammalian Wildlife Value
   /. Acute Toxicity
       The toxicity of 2,3,7,8-TCDD to mammals varies greatly both across mammalian species
   and within a given mammalian species. Large differences between mammalian species exist in the
   lethal dosages and toxic effects associated with acute doses of 2,3,7,8-TCDD. A difference of
   more than 8400 fold for LD*, values following single oral doses exists between guinea pigs (0.6-2
   /ig/kg) and hamsters (1157-5051 ugllug) (Harless et al., 1982; Kociba and Schwetz,  1982).
   Intraspecific differences in acute responses within a single species have also been observed. For
   example, LDjo values following oral exposure to 2,3,7,8-TCDD have varied from 182 to 2570
   /tg/kg in three different strains of mice. Rozman (1984) determined the intraperitoneal (i.p.) LD%
   dose of 2,3,7,8-TCDD to be 60 /ig/kg in rats. Mammalian LD^, values for 2,3,7,8-TCDD are
   summarized in Table 3-1.
       Acute toxic responses to 2,3,7,8-TCDD by mammals has been characterized by progressive
   loss of body weight, appetite suppression, and delayed lethality (Eisler, 1986). Rats treated with a
   single oral dose of 2,3,7,8-TCDD (5,  15, 25,  and 50 /tg/kg) have displayed a dose-related
   depression in food intake and body weight (Seefeld and Peterson, 1983). This "wasting
   syndrome" has been characterized in rats i.p. dosed with 60 /tg/kg of 2,3,7,8-TCDD (Rozman,
   1984), and rats provided a single oral dose of 50 /ig/kg (Seefeld et al.,  1984).
                      Table 3-1. Mammalian Acute Toxicity Values
Route
oral
oral
oral
oral
oral
Species
guinea pig
rat
Rhesus monkey
dog
mouse
LD,0 (/ig/kg)
0.6- 19
22-45
< 70
100-200
114-2570
                                                                              3-1

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                       Table 3-1.  Mammalian Acute Toxicity Values (Cont.)
Route
oral
oral
i.p.
oral
Species
rabbit
hamster
mink (newborn)
mink (adult)
LDM U/g/kg)
115
1157-5051
1'
4.2"
            Source: Eisier (1986), except for 'Autarch et al. (1988) and "Hochstein et al. (1988).
         The most sensitive wildlife mammalian species tested was the mink (Mustela visori). The i.p.
       j,, value for 2,3,7,8-TCDD was determined to be 1 /tg/kg in newborn mink following 12 days'
    exposure (Aulerich et al., 1988). The 28-day oral LDW was determined to be 4.2 fig/kg for adult
    mink (Hochstein et al., 1988).
    //. Chronic Tox/city
         No subchronic or chronic studies were obtained for mammalian wildlife species, however,
    chronic toxicity of 2,3,7,8 TCDD in wildlife species can be extrapolated from results of a number
    of subchronic and chronic studies using laboratory animals.
         Kociba et al. (1978) reported on a two-year toxicity and oncogeny study, using rats
    (Sprague-Dawley, 50 males and 50 females per group) administered dietary doses of 0, 0.001,
    0.01, and 0.1 /tg/kg/day. Hematological endpoints, urinary parameters, and gross and
    microscopic observations on tissues for tumors and tumor-like lesions were evaluated.  Animals
    given the high dose (0.1 /tg/kg/day) exhibited increased  mortality, decreased weight gain, slight
    depression of erythroid parameters, increased urinary excretion of porphyrins and delta-
    aminolevulinic acid and increased serum levels of certain enzymes. Increased tumor incidence and
    histopathologic or gross effects were observed in liver, lymphoid, lung, and vascular tissues of
    the high dose animals, and to a lesser extent in the mid-dose group. The liver was the organ that
    was most consistently affected in rats given 0.1  or 0.01  /tg/kg/day, exhibiting multiple
    hepatocellular degenerative, inflammatory, and necrotic  changes. Kociba et al. (1978) concluded
    that lifetime ingestion of 0.001 /tg/kg/day caused no effects considered to be of any lexicological
    significance. This study, therefore, reported a NOAEL of 0.001 /tg/kg/day, and a LOAEL of
    0.01 /tg/kg/day for liver effects in rats.
         Khera and Ruddick (1972) assessed the postnatal effect of prenatal exposure to 2,3,7,8-
    TCDD. Pregnant Wistar rats were given 0, 0.125, 0.25, 0.5, or  1.0 /tg/kg/day TCDD from days
    6 through 15 of gestation. Dose-related decreases  in the  average litter size and pup weight at birth
    were noted in all  but the 0.125 /tg/kg/day dose level. Survival of pups to 21  days was
    significantly reduced at 0.5 /tg/kg/day, and there were no surviving pups at 1.0 /tg/kg/day. In
    addition, decreases in the incidence of pregnancy and average litter size were noted in the f,
    generation of the  0.05 /tg/kg/day group. These results suggest a NOAEL of 0. 125 and a LOAEL
    of 0.25 /tg/kg/day for reproductive effects of TCDD on  Wistar rats.
         Murray et al. (1979) exposed three generations of Sprague-Dawley rats to dietary 2,3,7,8-
    TCDD. Rats were maintained  on diets equivalent to daily intake rates of 0, 0.001, 0.01, and 0.1
    /tg/kg/day for at least 90 days prior to gestation and throughout the gestation period. Fertility was
    significantly reduced among the rats  given 0.1 /tg/kg/day. At the 0.01 /tg/kg/day dose, no effect
    on fertility was observed among the fo rats, but a significant  reduction in fertility was observed
    among the ft and  f2 rats. No significant difference was observed between the fertility of the 0.001
    /tg/kg/day rats and the controls. Significantly decreased  litter sizes and gestational survival were
3-2

-------
noted among the f0 0.1 pg/kg/day group and the f, and fz rats receiving TCDD at 0.01 pg/kg/day.
Gestational survival was also significantly reduced among the 0.001 /ig/kg/day f2 generation, but
not in earlier or later generations. Significant decreases in postnatal body weight were observed
among the f2 and f3 litters but not the f, litter of the 0.01  /tg/kg/day group. Average body weight
of pups of rats given 0.1 /ig/kg/day, or any generation of the 0.01  /tg/kg/day group, were not
significantly different from those of control pups. The reproductive capacity of rats in the 0.001
/ig/kg/d group was not significantly affected in any generation, but it was reduced in the f, and f2
generations of the 0.01 ;ig/kg/day group. Therefore, a NOAEL of 0.001 ngfcgld and a LOAEL
of 0.01  /tg/kg/d for reproductive capacity of Sprague-Dawley rats were determined from this
study.
     Bowman et al. (1989, 1989a) reported impaired reproductive success of Rhesus  monkeys
exposed to 25 parts per trillion (ppt) (0.67 ng/kg/day) but not to 5 ppt (0.13 ng/kg/day) 2,3,7,8-
TCDD hi feed after 7 and 24 months. The exposures were discontinued after 4 years, and
breeding 10 months post-exposures for 4 years did not indicate reproductive impairment. The
offspring from these breeding experiments were evaluated for development and behavioral effects
(Bowman et al., 1989). While no significant effects of TCDD exposure were found on birth
weight, growth, or physical appearance of the offspring, results of  some behavioral tests,
including alterations in social behavior, were considered to be indicative of TCDD effects.  The
reproduction study of Bowman et al.  (1989a) provides clear evidence of a LOAEL at  0.67
ng/kg/day and a NOAEL at 0.13 hg/kg/day for reproductive effects of TCDD on Rhesus
monkeys.
     Studies  by Schantz et al. (1979), and Allen et al. (1979) also suggest that Rhesus monkeys
are more sensitive to 2,3,7,8-TCDD than rats. Decreases in fertility, increased abortions and
other toxic effects (e.g., alopecia, hyperkeratosis, weight loss, decreased hematocrit and white
blood cell count, and increased serum levels of SGTP) were noted at dietary levels of 50 ppt (1.5
ng/kg/day).
     The results of some of the studies described previously are summarized in Table 3-2.  The
study reported by Murray et al. (1979), in which three generations  of Sprague-Dawley rats were
exposed to 2,3,7,8-TCDD, was selected for use in developing the mammalian wildlife value
(WV). This study was selected because it consists of a multi-generational study that demonstrates
a dose-response to 2,3,7,8-TCDD exposure for reproductive effects. The reproduction study by
Bowman et al. (1989, 1989a) on Rhesus monkeys was not selected, although a lower  NOAEL
was determined in this  study, because the study by Murray et al.  (1979) incorporated  a multi-
generational exposure regime. In addition, the lack of comparative  toxicity endpoints for the mink
or river otter and the Rhesus monkey make it very difficult to estimate interspecies uncertainty
factors to apply if this study  were used.
                   Table 3-2. Summary of Chronic Mammalian Studies
Species
Rat
Rat
Rat
Rhesus
Monkey
LOAEL
(/sg/kg/day)
0.01
0.25
0.01
6.7 x 10-4
NOAEL
(/ig/kg/day)
0.001
0.125
0.001
1.3X10-4
Toxic Effect Observed
Liver effects
Reproductive
Reproductive
Reproductive
Reference
Kociba et al., 1978
Khera and Ruddick,
1972
Murray et al.,
1979
Bowman et al.,
1989, 1989a
                                                                                       3-3

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    ///.  Mammalian  Wildlife  Value Calculation
         In calculating a mammalian wildlife value, a species sensitivity factor (SSF) of 0.1 is used to
    reflect the uncertainty in extrapolating toxicity data from the rat to the mink and river otter. This
    SSF value is supported by the limited number of chronic toxicity studies available, the limited
    number of mammalian species for which chronic data is available, and the extreme sensitivity of
    mink among those mammalian species for which acute toxicity data is available.
    The wildlife equation and input parameters are presented below.
                    NOAEL (mammalian)
                   BAF3 (Trophic Level 3)
                   BAF4 (Trophic Level 4)
                                    SSF
                        0.001 //o/kg body weight/day
                        79,000 I/kg body weight
                        79,000 //kg body weight
                        0.1 (mink and otter)
    Body weights (WtJ, ingestion rates (FJ, and drinking rates (WJ for mink and river otter are
    presented in Table D-2 of the methodology document in Appendix D to 40 CFR 132 and shown
    below.
                                WtA (mink)
                                WtA (otter)
                                  FA (mink)
                                  FA (otter)
                                 WA (mink)
                                 WA (otter)
                           1.0kg
                           8.0kg
                           0.15 kg/day
                           0.9 kg/day
                           0.099 I/day
                           0.64 //day
    The equations and calculations of mammalian wildlife values are presented below.
     WV (mink)


     WV (mink)

     WV (mink)

     WV (otter)   =

     WV (otter)

     WV (otter)
                                             [NOAEL x SSF] x
                             ((1.0)(FA(mW,xBAF3)]

              _ (0.001 //g/kg/d x 0.1) 1.0 kg _
               0.099 */d + [(1.01(0.15 kg/d)(79,000 I/kg))

                           8.4 x 103pg/f

                       [NOAEL x SSF1 x WtA(omf,
                    l(0.5)(FA(omr| x BAF3) + (0.5)(FA(onw) x BAF4)]

                      (0.001 ;/g/kg/d xO.1) 8.0kg
0.64 */d +  [(0.51(0.90 kg/d x 79,000 I/kg)  +  l(0.5)(0.90 kg/d x 79,000 //kg)]

                            1.1 x 102pg/l
3-4

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The geometric mean of these two mammalian wildlife values results in
                        WV (mammalian)   =   e("nWVlmir*1 + klWV(etw"1/21

                        WV (mammalian)   =   e'"n -0084 " + h-01'P8"1/2'

                        WV (mammalian)   =   9.6 x 10'3 pg/f.
iv. Sensitivity Analysis for Mammalian Wildlife Value
     The values of the various parameters used to derive the mammalian wildlife value presented
above represent the most reasonable assumptions. The purpose of this section is to illustrate the
significance of these assumptions and the variability in the mammalian wildlife value if other
assumptions are made for the values of the various parameters from which the mammalian
wildlife value is derived. The intent of this section is to let the risk manager know, as much as
possible, the influence on the magnitude of the mammalian wildlife value of the assumptions
made in its derivation.
     In deriving the mammalian wildlife value for 2,3,7,8-TCDD, it was assumed that 100
percent of the mink diet was comprised of fish, although this may not necessarily be the case.
This assumption may lead to an overestimate of the 2,3,7,8-TCDD exposure for mink that are not
primarily feeding  on fish and aquatic invertebrates. As indicated in the Technical  Support
Document for Wildlife Criteria (Appendix to the  Preamble to 40 CFR 132), the fish  content of a
mink diet can vary from less than SO percent to the 100 percent assumed in the mink wildlife
value derivation presented above. If it were assumed only  50 percent of a mink's  diet was from
aquatic resources  and the remaining SO percent of the diet was uncontaminated, the estimated
2,3,7,8-TCDD exposure would be reduced by a factor of 2.  The resulting mink wildlife value
would be 1.7 x 10'2 pg/l, and the mammalian wildlife value would be 1.4 x 10*2 pg/l,  rather than
the mammalian wildlife value of 9.6 x 10~3 pg/l.
     As with all criterion derivations, there are uncertainties in assessing risk. The NOAEL
derived from Murray et al.(1979) for reproductive effects of 2,3,7,8-TCDD on rats concludes that
no adverse effects will be observed at that dose. However, a reevaluation of the Murray et al.
(1979)  data by Nisbet and Paxton (1982) using different statistical methods (i.e. pooling data from
different generations) indicated that both lower dose levels resulted in toxic effects, including
significant reductions in offspring survival indices, increases in liver and kidney weight of pups,
decreased thymus  weight of pups, decreased neonatal  weights, and increased  incidence of dilated
renal pelvis. Nisbet and Paxton (1982) concluded that 0.001  /ig/kg/day was a LOAEL and not a
NOAEL in the Murray et al. (1979) study. Another evaluation by Kimmel  (1988) considered the
Murray et al. (1979) data to be suggestive of a pattern of decreased offspring survival and
increased offspring renal pathology at 0.001 /ig/kg/day, even though the pooling of generations
by Nisbet and Paxton (1982) was considered to be biologically inappropriate. Assuming that
O.OQi Mg/kg/day is a LOAEL, and dividing this LOAEL by  a LOAEL to NOAEL uncertainty
factor of 10 results in a mammalian wildlife criterion value of 9.6 x  KT* pg/l instead of the
mammalian wildlife value of 9.6 x 10"3 pg/l.
     The mammalian assessment assumed that the mink is one of the most sensitive mammalian
species to the toxic effects of TCDD, and a SSF value of 0.1 was used to estimate a  mink/otter
NOAEL from the  rat NOAEL. A comparison of toxic effect levels between the Rhesus monkey
(Schantz et al.,  1979; Allen et al., 1979; Bowman et al., 1989, 1989a) and the rat (Murray et al.,
1979) suggests that the monkey may also  be quite sensitive. The mammalian wildlife value may
be calculated using the Rhesus monkey NOAEL of 0.13 ng/kg/day from the studies by Bowman
et al. (1989, 1989a) (a NOAEL value approximately 8 times lower than that determined for the
rat). Use of a SSF of 0.1 and the monkey reproductive NOAEL would result in a mammalian
wildlife value of 1.3 x 10"3 pg/l. However, in such an analysis, the use of a SSF of 0.1 to predict
a mink/otter NOAEL from the result of the Rhesus monkey study may be unduly  conservative,
                                                                                      3-5

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    given that the monkey and mink both appear to be extremely sensitive to the toxic effects of
    2,3,7,8-TCDD. Use of the Rhesus monkey NOAEL of 0.13 ng/kg/day, and an intermediate SSF
    of 0.3 would result in a mammalian wildlife value of 3.8 x 1C"3 pg/l compared to 9.6 x 10*3 pg/l.


III.  Calculation of Avian Wildlife Value
    /. Acute  Toxicity
         Single oral LDj,, values in avian species for 2,3,7,8-TCDD were reported by Eisler (1986).
    These LD values vary from 15 /tg/kg in Northern bobwhite quail (Colinus virginianus) to more
    than 810 pg/kg for the ringed turtle dove (Streptopelia risoria). Mallards (Anas platyrhynchos)
    are intermediate in sensitivity with an acute oral LDa value of more than 108 /tg/kg. Domestic
    chickens are relatively more sensitive to 2,3,7,8-TCDD than other avian species with an estimated
    oral LDjo range of 25 to 50 /tg/kg. The LD^ values for avian species are summarized in
    Table 3-3.
         Acute toxic responses to TCDD in birds are characterized by enlarged livers (turtle doves
    and domestic chickens), emaciation, vomiting, excessive drinking, central nervous system effects
    (bobwhite quail), and signs of chick edema disease (chickens) (Eisler, 1986).

                             Table 3-3. Avian Acute Toxicity Values
Species
Northern bobwhite quail (Colinus virginianus)
Ringed turtle-dove (Streptopelia risoria)
Mallard (Anas platyrhynchos)
Domestic chicken (Gallus domesticus)
LDM 0/g/kg)
15
> 810
> 108
25-50
         Source: Eisler (1986).
    //'. Chronic Toxicity
         Environmental mixtures of halogenated aromatic hydrocarbons have been implicated in a
    number of adverse effects including reproductive failure in avian species (Gilbertson et al., 1991).
    In such field studies, the observation of reduced reproduction has been correlated to 2,3,7,8-
    TCDD equivalents; however, the dose-response relationship specific to 2,3,7,8-TCDD itself
    cannot be discerned from the effects of other contaminants. Most of the laboratory research
    directed at the determination of dose-response relationships with TCDD has been based on
    mammalian species, with very little attention given to avian species.
         The research of Nosek et al. (1992, 1992a, and 1993) represents the only comprehensive
    laboratory investigation of the subchronic toxicity and toxicokinetics of 2,3,7,8-TCDD among
    avian species. Ring-necked pheasants (Phosianus colchicus) were dosed weekly, intraperitoneally,
    for 10 weeks at an equivalent rate of 0.14, 0.014 and 0.0014 /tg/kg/day. Egg production was
    significantly reduced among pheasants from the 0.14 ng/kg/day group, but not in pheasants from
    the two lowest dose groups when compared to controls. In addition, the 0.14 /xg/kg/day dose was
    associated with a significant increase in mortality of embryos from the fertilized eggs of those
    hens. The LOAEL for embryo mortality and fertility was 0.14 pg/kg/day; therefore, the NOAEL
    determined from this study is 0.014 /tg/kg/day.
         The reproductive effect NOAEL for 2,3,7,8-TCDD determined from the Nosek et al. (1992,
    1992a, and 1993) studies will be used in calculating the avian wildlife value. The data generated
    from this study show a clear dose-response w.ih a meaningful endpoint and are based on
    exposures lasting more than 28 days. This study is based on an i.p. injection study rather than an
    oral route of administration. It is generally acknowledged that i.p. and oral routes of exposure are
3-6

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similar because in both instances the chemical is absorbed first by important internal organs such
as the liver, thereby permitting first-pass metabolism. Use of the i.p. dose levels assumes that
2,3,7,8-TCDD bioavailability and absorption from the gastrointestinal tract and the abdominal
cavity are not significantly different. To the extent that an i.p. exposure results in higher or lower
2,3,7,8-TCDD absorption than that associated with an oral exposure, the hazards to avian wildlife
may be over- or under-estimated.

i/7. Avian Wildlife Value Calculation
     In calculating a wildlife value for kingfisher, osprey, and eagle, a SSF of 0.1 is used
because of the uncertainty in extrapolating data across species given the paucity of chronic
toxicity data. In  addition, a comparison of results from in ovo studies indicate the chicken (Allred
and Strange, 1977) is approximately five times more sensitive than the pheasant (Nosek et al.
1992a). Also, a comparison of single-dose LDjo values suggests that although the pheasant is
among the more sensitive species tested, there are other birds more susceptible to 2,3,7,8-TCDD
intoxication (Eisler, 1986; Nosek et al.  1993).
The wildlife equation and input parameters are presented below.
                      NOAEL (avian)
              BAF3 (Trophic Level 3)
              BAF4 (Trophic Level 4)
                                SSF
0.014//Q/kg/day
79,000 f/kg body weight
79,000  /kg body weight
0.1 (kingfisher, osprey and eagle)
Values for body weights (WtyJ, ingestion rate (F^, and drinking rate (W^ for kingfisher, osprey,
and eagle are presented in Table D-2 of the methodology document (Appendix D to 40 CFR  132)
and shown below.
                          WtA (kingfisher)
                             WtA (osprey)
                              WtA (eagle)
                           FA (kingfisher)
                              FA (osprey)
                                FA (eagle)
                          WA (kingfisher)
                             WA (osprey)
                              WA (eagle)
     0.15kg
     1.5kg
     4.5kg
     0.075 kg/day
     0.3 kg/day
     0.5 kg/day
     0.017 f/day
     0.077 //day
     0.16  I/day
                                                                                       3-7

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    Calculations of avian wildlife values are summarized below.

                                              [NOAEl x SSF] x
                WV (kingfisher)
                                          W >-	t^t	4- f/1 OHF       x BAF 11


                                      	(0.014 j/g/kg/d x 0.1) 0.15 kg	
                WV (kingfisher)    =    0.017 //d + [(1.0)(0.075 kg/d x 79,000 //kg)]


                WV (kingfisher)    =                 3.5 x 10 2 pg//
                   WV (osprey)



                   WV (osprey)

                   WV (osprey)



                    WV (eagle)
                                               [NOAEL x SSF] x
                                                              x BAF3

                                  	(0.014//g/kg/dx0.1) 1.5kg
                                   0.077 //d + K1.0H0.3 kg/d x 79,000 //kg)]

                                                8.9 x 102pg//


                                           [NOAEL x SSF] x WtA(MB..,
                                        WA(MoW + [<1.0)(FA(M8)a) x BAF4)]

                                          (0.014//g/kg/d x 0.1) 4.5 kg
                WV (eagle)    =     0.16 //d +  [(1.01(0.5 kg/d x 79,000 //kg)]

                WV (eagle)    =                  1.6x101pg//

The geometric mean of these three avian wildlife values results in
                   WV (avian)
                                         e
                                          Illn WVdungfntwr) + In WV(epry) + In WVIglll/3)
                   WV (avian)   =   ellln '03B P8// *
                   WV (avian)   =   7.9 x 10'2 pg/l.
                                                               io.ipg/iw)
    iv. Sensitivity Analysis of Avian Wildlife Value
         The values of the various parameters used to derive the avian wildlife value presented above
    represent the most reasonable assumptions. The purpose of this section is to illustrate the
    significance of these assumptions and the variability in the avian wildlife value if other
    assumptions are made for the values of the various parameters from which the avian wildlife
    value is derived. The intent of this section to let the risk manager know, as much as possible, the
    influence on the magnitude of the avian wildlife value of the assumptions made in its derivation.
         The lack of chronic toxicity data for avian species other than pheasants results hi some
    uncertainty associated with the development of the avian wildlife value. Based on the single
    exposure acute toxicity data, it could be argued that species such as bobwhite quail, pheasant, and
    chicken may be among the most sensitive avian species to 2,3,7,8-TCDD.  If this were indeed the
    case, an intermediate SSF value of 0.3 rather than 0.1 may be appropriate. Use of this uncertainty
    factor would result in an avian wildlife value of 2.4 x Ifr'pg/f instead of 7.9  x 10~2 pg/t.
         A subchronic to chronic uncertainty factor in the avian wildlife value calculation was not
    used, although the study duration was only 10 weeks, because the reported toxicokinetics of
    2,3,7,8-TCDD in laying pheasant hens suggest that a significant portion of the dose is excreted
    (Nosek et al., 1992). However,  if a  '0-fold subchronic to chronic uncertainty factor were used,
    the avian criterion would be 7.9 x 10"3pg/f. If the  10-fold subchronic to chronic uncertainty
3-8

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    factor were used in combination with a SSF value of 0.3, the avian wildlife criterion would be
    2.4 x  10-2pg/f rather than 7.9 x 10* pglt.
        The derivation of an avian wildlife value is based on the assumption that 100 percent of an
    eagle's diet is composed of fish. A study by Kozie and Anderson (1991) suggests that fish
    comprise 97 percent of Lake Superior eagle diets and mammals and birds each comprise 1.5
    percent of eagle diets. Assuming the metabolizable energy in fish is approximately 1  kcal/g
    (Palmer,  1988; and Stalmaster and Gessaman, 1982) and the typical eagle consumes about 500 g
    of fish per day (Technical Support Document for Wildlife Criteria,  Appendix to the Preamble to
    40 CFR 132), in eagle has a daily energy requirement of 500 kcal/day. The energy content for
    birds is 2 kcal/g (a value derived for mallards; Stalmaster and Gessaman, 1982). Applying the
    conservative assumptions that the bioaccumulation in mammals would be equivalent to that in
    Trophic Level 4 fish and the caloric value would be the same for mammals and fish, an eagle diet
    consisting of 1.5 percent fish-eating birds and 98.5 percent fish would result in a daily intake of
    approximately 7.4 g of bird and 480 g of fish to meet the daily energy requirement of 500
    kcal/day. Braune and Nordstrom (1989) have reported that 2,3,7,8-TCDD bioaccummulates in
    Lake Ontario herring gulls at a level approximately 30 times higher than that observed in alewife.
    Therefore, dietary exposure to eagles of 2,3,7,8-TCDD would be higher if piscivorous birds
    comprise a portion of their diets. The 2,3,7,8-TCDD exposure to eagles eating 7.4 g of
    piscivorous birds a day would be approximately 1.4  times higher than an exposure associated with
    a 100  percent fish diet. Such an analysis would result in a bald eagle wildlife value of 1.1 x  10*1
    pg/f,  and an avian wildlife value of 7.1 x 10"2 pg/t compared to 7.9 x 10"2  pglt.


IV. Great Lakes Wildlife Criterion
        The Great Lakes Wildlife Criterion for 2,3,7,8-TCDD is determined by the lower of the
    mammalian wildlife value (9.6 x 10~3 pg/l)  and the avian wildlife value (7.9 x 10"2 pg/l)- 'The
    mammalian wildlife value was determined to be approximately one  order of magnitude smaller
    than the avian wildlife value. Therefore, the Great Lakes Wildlife Criterion for 2,3,7,8-TCDD is
    9.6 x  lO"3 pg/f.


    /. Discuss/on of  Uncertainties
        Wildlife populations inhabiting the Great Lakes Basin would not be impacted from the intake
    of drinking water or aquatic prey taken from surface water containing 2,3,7,8-TCDD in
    concentrations of 9.6  x 10"3 pg/t, based on  the uncertainty factors used to account for data gaps
    and the variability in the toxicity and exposure parameters inherent  in the 2,3,7,8-TCDD risk
    assessment. Criteria for other ecoregions may require an analysis of different wildlife species with
    different diets and body masses. In addition, the bioaccumulation factors in this analysis were
    based  on an analysis for the Great Lakes, and different bioaccumulation factors may be more
    appropriate for other waterbodies.
        Finally, generic  assumptions were made in assessing the hazards of 2,3,7,8-TCDD to
    wildlife populations through the use of LOAELs and HOAELs for reproduction arid development.
    The use of these levels assumes no hazards  to wildlife populations would result from the direct
    exposure of individuals to 2,3,7,8-TCDD. However, it could be  argued that some increase in
    density independent mortality, or decrease in density independent reproductive success, which
    could be attributable to 2,3,7,8-TCDD exposure could be incurred without impacting the
    population dynamics of a species. In general,  well-validated population models do not yet exist
    for the species analyzed, and it  is difficult to estimate the extent of mortality or reproductive
    failure that could be incurred. In addition, the interaction of additional chemical as well as non-
    chemical stressors on wildlife population responses is also poorly resolved at this time.
                                                                                           3-9

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V.  References
Allen, J.R., D.A. Banotti, L.K. Lambrecht, and J.P. Van Miller. 1979. Reproductive effects of halogenated aromatic
     hydrocarbons on nonhuman primates. Ann. NY Acad. Sci. 320:419-425.
Allred, P.M. and J.R. Strange. 1977. The effects of 2,4,5-tricnloropnenoxyacetic acid and 2,3,7,8-tetrachlorodibenzo-/>-
     dioxin on developing chick embryos. Arch. Environ. Contain. Toxicol. 5:483-489.
Aulerkh, RJ., S.J. Bursian and A.C. Napolitano.  1988. Biological effects of epidermal growth factor and 2,3,7,8-
     tetrachlorodibenzo-f-dioxin on developmental parameters of neonatal mink. Arch. Environ. Contain. Toxicol. 17:27-
     31.
Bowman, R.E., S.L. Schantz, M.L. Gross, S. Ferguson. 1989. Behavioral effects in monkeys exposed to 2,3,7,8-
     tetrachlorodibenzo-p-dioxin transmitted maternally during gestation and for four months of nursing. Chemosphere 18(1-
     6):235-242.
Bowman, R.E., S.L. Schantz, N.C.A. Weerasinghe, M.L. Gross, D.A. Barsotti. 1989a. Chronic dietary intake of 2,3,7,8-
     tetrachlorodibenzo-p-dioxin (TCDD) at 5 and 25 parts per trillion in the monkey: TCDD kinetics and dose-effect
     estimate of reproductive toxicity. Chemosphere 18(l-6):243-252.
Braune, B. M. and R. J. Norstrom. 1989. Dynamics of organochlorine compounds in herring gulls: HI.  Tissue distribution
     and bioaccumulation in Lake Ontario gulls. Environ. Toxicol. Chem. 8:957-968.
Efeter, R. 1986. Dioxin  Hazards to Fish, Wildlife and Invertebrates: A Synoptic Review. U.S.  Fish Wild. Serv. Biol.  Rep.
     85(1.8). 37 pp.
Gilbertson, M., T. Kubiak, J. Ludwig, G. Fox. 1991. Great Lakes embryo mortality, edema, and deformaties syndrome
     (GLEMEDS) in colonial fish-eating birds: Similarity to chick edema disease. J. Toxicol. Environ. Health 33:455-520.
Harkss, R.L., E.O. Oswald, R.G. Lewis, A.E. Dupuy Jr., D.D. McDaniel, and H. Tai. 1982. Determination of 2,3,7,8-
     tetrachlorodibenzo-/>-dioxin in fresh water fish. Chemosphere 11:193-198.
Hochstein, J.R.,  R.J. Aulerich and S.J. Bursian. 1988. Acute toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in mink.
     Arch. Environ. Con tarn. Toxicol.  17:33-37.
Khera, K.S., and J.A. Ruddick.  1972. Polychlorodibenzo-/>-dioxins: Perinatal effects and the dominant lethal test in Wistar
     rats. In E.H. Blair,  ed. Chlorodioxins - Origin and Fate Advances in Chemistry Series 123. Amer. Chem. Soc.,
     Washington, D.C.
Kimmel, G. L. 1988. Appendix C: Reproductive and developmental toxicity of 2,3,7,8-TCDD in A Cancer Risk-Specific
     Dose Estimate for 2,3,7,8-TCDD. Review draft. EPA/600/6-88/007Aa.
Kociba, R J., D.G. Keyes, J.E. Beyer, R.M. Carreon, C.E. Wade, D.A. Dittenber, R.P. Kalnins, L.E. Frauson, C.N.
     Park, S.D. Barnard, R.A. Hummel, and C.G.  Humiston. 1978. Results of a two-year chronic toxicity and oncogenicity
     study of 2,3,7,8-tetrachlorodibenzo-/>-dioxin in rats. Toxicol.  Applied Pharmacol. 46:279-303.
Kociba RJ. and B.A. Schwetz. 1982. Toxicity of 2,3,7,8-tetrachlorobenzo-/>-dioxin (TCDD). Drug Metab.  Rev. 13:387-
     406.
Kozie K.D. and R.K. Anderson. 1991. Productivity, diet, and environmental contaminants in bald eagles nesting near the
     Wisconsin shoreline of Lake Superior. Arch. Environ. Contain. Toxicol. 20:41-48.
Murray, FJ., F.A. Smith, K.D. Nitschke, C.G. Huniston, R.J. Kociba and B.A. Schwetz. 1979. Three-generation
     reproduction study of rats given 2,3,7,8-tetrachlorodobenzo-^-dioxin (TCDD) in the diet. Toxicol. Applied Pharmacol.
     50:241-252.
Nisbet, I.C.T. and M.B. Paxton. 1982. Statistical aspects of three-generation studies of the reproductive toxicity of TCDD
     and 2,4,5-7.  The American Statistician 36(3):29O-298.
Nosek, J. A., J. R. Sullivan, T. E. Amundson, S. R. Craven, L.  M. Miller, A.  G. Fitzpatrick, M. E. Cook and R. E.
     Peterson. In Press.  Embryotoxicity of 2,3,7,8-tctrachlorodibcnzo-p-dioxin in ring-necked pheasants. Environ. Toxicol.
     and Chem.
Nosek, J. A., J. R. Sullivan, S. S. Hurley, J. R. Olson, S. R. Craven, and R. E. Peterson. 1992. Metabolism and
     disposition of 2,3,7,8-tetrachlorodibenzo-/>-dioxin in ring-necked pheasant hens, chicks, and eggs. Journal of
     Toxicology and Environmental Health 35:153-164.
Nosek, J.A., J.R. Sullivan, S.S. Hurley, S.R.  Craven, and R.E. Peterson. 1992a. Toxicity and reproductive effects of
     2,3,7,8-tetrachlorodibenzo-/>-dioxin toxicity in ring-necked  pheasant hens. Journal of Toxicology and Environmental
     Health 35:187-198.
Palmer, R.S., Editor. 1988. Handbook of North American Birds: Vol. 4. Yale University Press. 433 pp.
Rozman, K. 1984. Separation of wasting syndrome and lethality caused by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol.
     Letters. 22:279-285.
Schantz, S.L., D.A. Barsotti, J.R. Allen. 1979. Toxicological effects produced in nonhuman primates chronically exposed
     to fifty parts per trillion 2,3,7,8-tetrachlorodibenzo-/>-dioxin (TCDD). Toxicol. Appl. Pharmacol. 48:(2)A180 (Abstract
     No. 360).
Seefeld, M.D.,  S.W. Corbett, R.E. Keesey and R.E. Peterson.  1984. Characterization of the wasting syndrome in rats
     treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. Appl. Pharmacol. 73:311-322.
Seefeld, M.D. and R.E.  Peterson. 1983. 2,3,7,8-Tetrachlorodibenzo-/>-dioxin-induced weight loss, pp 405-413 in Tucker,
     R.E.  et al., eds. Human and Environmental Risks of Chlorinated Dioxins and Related Compounds. Plenum, New
     York.
Stalmaster M.V. and J.A. Gessaman.  1982. Food consumption and energy requirements of captive bald eagles. J. Wildl.
     Managem. 46:646-654.
3-10

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CHAPTER  4
Tier  I  Wildlife Criteria  for
Polychlorinated Biphenyls  (PCBs)	


                               Contents
I.   Literature Review	   4-1
n.  Calculation of Mammalian Wildlife Value	   4-1
    i.  Acute Toxicity	   4-1
    ii.  Chronic Toxicity	   4-1
    iii.  Mammalian Wildlife Value Calculation	   4-5
    iv.  Sensitivity Analysis for Mammalian Wildlife Value  	   4-7
m.  Calculation of Avian Wildlife Value  	   4-7
    i.  Acute Toxicity	   4-7
    ii.  Chronic Toxicity	   4-8
    iii.  Avian Wildlife Value Calculation  	   4-11
    iv.  Sensitivity Analysis of Avian Wildlife Value 	   4-12
IV.  Great Lakes Wildlife Criterion	   4-13
    i.  Discussion of Uncertainties 	   4-13
V.  References  	   4-14

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Tier I  Wildlife  Criteria  for
Polychlorinated  Biphenyls  (PCBs)
1.  Literature Review
        A review of mammalian and avian toxicity data for polychlorinated biphenyls was based on
   literature received through computer-based (CAS and BIOSES) as well as manual searches. A
   total of 41 references were screened; those references which were reviewed in detail are cited in
   Section V and primarily include those that contain dose-response data.


II. Calculation  of Mammalian Wildlife Value
   /. Acute Toxic/ty
        Three primary effects of PCB exposure on terrestrial wildlife are mortality, decreased
   reproductive success, and behavioral modifications. Mink (Mustela visori) appear to be among the
   more sensitive species to the toxic effects of PCBs (Gillette et al., 1987). Single oral doses of
   PCBs administered to mink have produced LDjo values of 750 mg/kg for Aroclor 1221 and 4000
   mg/kg for Aroclor 1254 (Aulerich and Ringer,  1977; Ringer,  1983). Diets containing PCBs at
   6.7 mg (Aroclor 1254)/kg to 8.6 mg (Aroclor 1242)/kg fresh weight have caused 50 percent
   mortality among mink over a 9-month period (Ringer, 1983). The reasons for mink sensitivity to
   PCBs are unknown, but interspecific variability in sensitivity to PCBs is  common, even among
   closely-related species. For example, the acute toxicity (i.e., LCjo) of Aroclor  1242 has been
   demonstrated to be lower among European ferrets (Mustelaputorius furo) (LCX> 20 mg/kg)
   than among mink  (LC = 8.6 mg/kg) (Eisler, 1986). Age, dietary composition, season, and year
   have had little effect on the outcome of the acute toxicity tests. Acute toxicity values for mink fed
   a diet containing a mixture of PCBs produced 28- and 35-day LDjo values of 79 and 48.5 ppm,
   respectively (Hornshaw et al., 1986). Acute toxicity values for mink fed  a diet consisting of cattle
   that had been fed Aroclor 1254 were 47 and 31.5 ppm, respectively for 28-day and 35-day LC^,
   tests (Aulerich et al., 1986). In a longer term study, high daily intake of PCBs (Clophen A-60,
   equivalent to Aroclor 1060) fed to female mink  for 51 days caused 100 percent mortality at 6.1
   mg/day and 40 percent mortality at 2.0 mg/day  (den Boer, 1984). Table  4-1 provides a summary
   of values of acute mammalian toxicity to PCBs.


   ii. Chronic Toxicity
        Numerous studies (Ringer et al., 1972; Platonow and Karstad, 1973; Jensen et al. 1977;
   Aulerich and Ringer, 1977; U.S. EPA, 1980; Bleavins et al. 1980) have  demonstrated that mink
   are among the most sensitive mammalian species to the toxic effects of PCBs, with some PCB
   congeners being more toxic than others. The main chronic effect that has been documented as a
   result of dietary exposure to PCBs has been decreased reproductive success, as evidenced by
   reduced  whelping  rates, fetal death,  and reduced growth among the young.
                                                                                   4-1

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                          Table 4-1. Mammalian Acute Toxicfty Values
PCB Congener
1221
1242
1254
1260
Route
oral
oral
dermal
i.p.
diet
diet
oral
oral
dermal
i.p.
metabolized diet
diet
diet
diet
diet
diet
oral
oral
i.p.
oral
dermal
Species
rat
mink
rabbit
mink
mink
ferret
rat
mink
rabbit
mink
mink
mink
mouse
rat
raccoon
rabbit
rat
mink
mink
rat
rabbit
LOW (g/kg)
1 -4
0.75 - 1
4
0.5 - 0.75
0.0086
0.02
0.8 - 8.7
3
8.7
1
0.0032 - 0.0047*
0.0067
> 0.1 - > 0.25
> 0.075
> 0.05
> 0.01
0.5- 1.4
4
1.25-2.25
1.3- 10
10
            Source: Eisler (1986), except for 'Aulerich et al. (1986).

         Bleavins et al. (1980) investigated the effects of dietary exposure to Aroclors 1016 and 1242
    on mink and ferrets. Mink were fed a diet supplemented with either 0, 5,  10, 20, or 40 ppm
    Aroclor 1242 or 20 ppm Aroclor 1016. The ferrets were fed a diet supplemented with either 0, or
    20 ppm Aroclor 1242 or 20 ppm Aroclor 1016 for 9 months. Aroclor  1242 produced 100 percent
    mortality in all adult mink fed diets at the 20 ppm and 40 ppm levels.  Only one female and no
    males died when fed a diet containing 5 ppm Aroclor 1242. No mortality was noted among mink
    fed diets containing 20 ppm Aroclor 1016. Mink fed Aroclor 1242 at 5 and 10 ppm failed to
    reproduce, while Aroclor 1016 reduced but did not completely eliminate reproduction. In contrast
    to these results, no mortality attributed to the PCBs  was observed among the ferrets. Ferrets fed
    the Aroclor 1242 diet did not whelp, but reproductive performance among the female ferrets fed
    Aroclor 1016 was not significantly different from that of the control females. In addition,
    reproductive parameters, kit growth, and adult and kit mortality were not significantly affected in
    the ferrets fed the Aroclor 1016 diet. Using a mink body weight of 1 kg, and a food consumption
    rate of C.15 kg/day, provided in the methodology document (Appendix D to 40 CFR 132), the
    results from this study suggest a mink reproductive LOAEL of 0.75 mg/kg/day (5 ppm) for
    Aroclor 1242 and 3.0 mg/kg/day (20 ppm) for Aroclor 1016. The body weights and food
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consumption rates of ferrets are virtually identical to mink (Ringer et al., 1981). Using a ferret
body weight of 1 kg and a food consumption rate of 0.15 kg/day, the LOAEL for reproductive
effects for Aroclor 1242 and the NOAEL for reproductive effects for Aroclor 1016 are 3.0
20.096 mg/mg/kg/day.
     Linzey (1988) evaluated reproductive success and growth among white footed mice
(Peromyscus leucopus) chronically exposed to Aroclor 1254 in the diet at a level of 10 ppm.
PCB-treated second generation mice exhibited poor reproductive success in comparison with
second generation controls and the parental generation. This was evidenced by reduced
reproductive organ weights, drastically reduced number of litters, and survival among ie young
of the  second generation treated group. Poor growth among the second generation PCB-treated
litter was also observed, with increasing differences in body weights becoming apparent over time
when compared to controls. Using a mouse body weight of 32 grams, and an ingestion rate  of 4.9
g/day (U.S. EPA,  1988), the dietary PCB exposure associated with reproductive effects was
calculated to be 1.53 mg/kg/day.
     According to Platonow and Karstad (1973) and Hornshaw et al. (1983), reproductive
impairment occurs in mink at even lower concentrations  when the PCBs fed to the mink have first
been metabolized by another species. Platonow  and Karstad (1973) orally dosed Aroclor 1254 to
Jersey cows, and fed the resulting contaminated beef to mink over 160 days at 0.64 and 3.57 ppm
total PCBs.  At a dietary concentration 3.57 ppm total PCBs, no live kits were produced and all
adult mink died before the end of the experiment. At 0.64 ppm total PCBs in the diet, 2 of  14
adult mink died before the end of the experiment and only 1 of 12 mink produced kits. All 3 of
the kits died during the first day after birth.  Based  on these findings the LOAEL for successful
reproduction was 0.64 ppm. Based on the mink body weight and food consumption rate presented
above, the LOAEL was calculated as 0.0% mg/kg/day for reproductive effects of total PCBs.
     Hornshaw et al. (1983) fed Great Lakes fish or fish products to mink for up to 290 days.
Dietary concentrations of PCB residues were determined to range from 0.21 to 1.50 ppm. Only
mink fed PCBs at concentrations of 0.21  ppm had reproduction and kit survival similar to the
controls. Mink fed a diet containing 0.48 ppm of PCB residues had inferior reproductive
performance and/or kit survival when compared to controls. These findings suggest a NOAEL of
0.21 ppm and  a LOAEL of 0.48 ppm. Using the mink body weight and food ingestion rate
presented above, the NOAEL was calculated to be  0.032 mg/kg/day, and the LOAEL was 0.072
mg/kg/day for reproductive performance  and kit survival. Hornshaw et al. (1983) observed that
the toxicity of PCBs was greater when derived from Great Lakes fish than in previous studies
using comparable levels of technical  grade PCBs. However, concentrations of other toxicants
potentially present in the Great Lakes fish were not measured.
     Fetotoxicity and reproductive failure have  also been reported for mink following direct
dietary exposure to low levels of certain PCB congeners. Wren et al. (1987) fed adult ranch-bred
mink diets containing either 0 or 1.0 ppm Aroclor  1254, 1.0 ppm methylmercury, a combination
of 1.0  ppm Aroclor 1254 and 1.0 ppm methylmercury, or a combination of 0.5 ppm Aroclor
1254 and 0.5 ppm methylmercury. Fertility of adult male mink, percentage of females whelped,
or number of kits born per female were not affected by the treatments, but the growth rate of the
kits nursed by the mothers exposed to 1.0 ppm Aroclor 1254 (0.15 mg/kg/day) was significantly
reduced.
     In a sub-chronic study Jensen et al.  (1977) dosed mink with PCBs (congeners not reported)
in the feed at concentrations of 0.05, 3.3, and 11 ppm for 66 days.  Complete reproductive failure
was observed among the 11 ppm group, with reduced number of implantation sites and no kits
born. The frequencies of mated and pregnant females did not differ significantly between the 0.05
ppm group and the 3.3 ppm dose group. At 3.3 ppm, however,, the frequency of delivering
females, and the number of kits born per female were significantly smaller than at 0.05 ppm. In
addition, the number of stillbirths at 3.3 ppm was greater and  the average body weight of the
young  was smaller than among the controls.  From these results, a LOAEL for reproduction  can
be inferred of 3.3 ppm, and a NOAEL of 0.05 ppm. Using the mink body weight and ingestion
                                                                                      4-3

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    rates presented previously, the LOAEL is calculated as O.S mg/kg/day, and the NOAEL is 0.008
    mg/kg/day.
         Aulerich and Ringer (1977) exposed mink to dietary Aroclor 1254 at 0, 5, and 10 ppm over
    a 9-month period. All of the mink fed PCB-supplemented diets failed to produce offspring. In a
    subsequent experiment, mink were provided diets containing 2 ppm Aroclor 1016, 1221, 1242, or
    1254, and monitored over 297 days. Aroclor 1254 was the only PCB mixture that had an adverse
    effect on reproduction. Two of the 7 females whelped and 1 live, underweight kit was produced.
    Based on these studies, a LOAEL for reproductive success of 2 ppm Aroclor 1254 can be
    inferred. Using the mink body weight and food consumption rates presented above, a LOAEL
    was calculated to be 0.3 mg/kg/d for reproductive effects of Aroclor 1254.
         Aulerich and Ringer (U.S. EPA, 1980) investigated the effects of Aroclor 1016 on
    reproduction, growth, and survival of mink. Mink were fed diets that contained 0, 2, 10, and 25
    ppm Aroclor 1016 for up to 18 months. No marked hematological changes or clinical signs of
    PCB poisoning were observed in even the highest dose group; however, increased heart and
    decreased kidney weights were noted in some animals, but these were not consistent among the
    treated groups. Reproduction was not adversely affected, but reduced 4-week weights were
    observed among kits nursed by females fed the 25 ppm PCB supplemented diet, and excessive kit
    mortality between birth and 4 weeks was noted among most of the groups provided with PCB
    supplemented diets. The authors attributed these adverse effects to quantitative or qualitative
    impacts of PCBs on lactation. From these results, a LOAEL of 2 ppm  for kit survival  and growth
    can be inferred. Using the mink body weight and feeding rate presented above, this LOAEL is
    equivalent to 0.3 mg/kg/day.
         Aulerich et al.  (1985) fed Aroclor 1254 and three hexachlorobiphenyl congeners
    (2,4,5,2',4',5'- [245 HCB]; 2,3,6,2',3',6'- [236 HCB]; and 3,4,5,3',4',5'- [345 HCB]) to adult
    female mink for 12.5 weeks at concentrations ranging from 0.1 ppm to 5.0 ppm in the diet (each
    congener was not given at each dose level).  Concentrations of 5 and 2.5 ppm of 245 HCB or 236
    HCB had no significant effect on the number of females that whelped or the litter size per female
    whelped. Only 1 out of 10 females whelped and no live kits were produced at 2.5 ppm Aroclor
    1254 in the diet. At 0.5 ppm  345 HCB in the diet, all animals died after 29 to 72 days exposure.
    At 0.1  ppm 345 HCB in the diet, 50 percent mortality was observed before the end of the
    experiment and none of the 8 females whelped. Based on the results of Aulerich et al.  (1985), a
    LOAEL for survival and for reproductive effects of 0.1 ppm 345 HCB can be inferred. Using the
    body weight and food ingestion rate provided above, mis LOAEL is equivalent to 0.015
    mg/kg/day for survival and reproductive effects of 345 HCB. The LOAEL from this study for
    reproductive effects of Aroclor 1254 is 0.375 mg/kg/day.
        den Boer (1984) investigated reproductive effects of dietary exposure to PCBs originating
    from fish livers and  Clophen  A-60 (equivalent to Aroclor 1260) during 400 days. Mink were
    maintained on feed contaminated with total PCBs at levels equivalent to 25.2 /ig/kg/day. No
    mortality was observed among the dosed groups; however, a significant reduction in females
    whelping was observed among the exposed mink.
        The various toxicity values derived from the studies that were discussed previously are
    summarized  in Table 4-2. An evaluation of these studies suggests that the LOAEL of 0.3 mg/kg/d
    for reproductive effects of Aroclor 1254, from the study of Aulerich and Ringer (1977), is the
    most appropriate daily dose rate to use in calculating a mammalian wildlife value  (WV) for total
    PCBs. The LOAEL values for mink developed for HCBs in the Aulerich et al. (1985) study are
    lower than the LOAEL for Aroclor 1254; however, they cannot be used for  criteria development
    because of a lack of dose-response data. Furthermore, use of the LOAEL for HCB would be
    based on the unreasonable assumption that all PCBs discharged into the environment are HCBs,
    or that all the discharged PCBs would be totally converted to 3,4,5-HCB. The LOAELs derived
    using metabolized PCBs (Platonow and Karstad, 1973; Hornshaw et al., 1983) are not appropriate
    for criteria development, in part because possible contamination of feed by additional
    environmental contaminants was not investigated. Therefore, the results of Aulerich and Ringer
4-4

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(1977) were considered to more properly reflect the toxicity of total PCBs to which mammalian
wildlife species are exposed.


///.  Mammalian Wildlife Value Calculation
     The LOAEL derived from Aulerich and Ringer (1977) of 0.3 mg/kg/d was based on a 297-
day feeding study. This LOAEL was divided by a subchronic to chronic uncertainty factor of 10,
resulting in an adjusted LOAEL of 0.03  mg/kg/d. This uncertainty factor was used based on the
results of the study by den Boer (1984) in which a significant reduction in females whelping was
observed in mink exposed to PCB-treated feed at a concentration of 0.025 mg/kg/day for 400
days, which corresponds to an intake rate approximately 10 times lower than the LOAEL derived
from Aulerich  and Ringer (1977). As discussed by den Boer  (1984), the results of these two
studies illustrate that the total amount of PCB intake, rather than the daily dose, is critical in
assessing adverse effects. Lower dietary PCB concentrations  can cause significant adverse effects
with a sufficiently long exposure duration.
     A NOAEL for reproductive effects in mink from total PCBs was determined by  dividing the
LOAEL by a LOAEL to NOAEL uncertainty factor of 10. Thus, the NOAEL used in calculating
mammalian wildlife values was 3.0 /tg/kg/d.
                   Table 4-2. Summary of Chronic Mammalian PCB Studies
Species
Mouse
Mink
Ferret
Mink
Mink
Mink
Mink
Mink
Mink
Mink
Mink
LOAEL
(mg/kg/day)
1.53
0.75
3.0
3.0

0.096
0.072
0.5
0.15
0.3



0.3
0.375


0.015
0.025
NOAEL
(mg/kg/day)




3.0

0.032



0.3
0.3
0.3


0.375
0.375


PCB Congener
Aroclor-1254
Aroclor-1242
Aroclor-1016
Aroclor-1 242
Aroclor-1016
Metabolized
Aroclor-1254
Metabolized
total PCBs
Unreported
PCBs
Aroclor-1254
Aroclor 1254
Aroclor-1016
Aroclor- 1021
Aroclor-1242
Aroclor-1016
Aroclor-1 254
245 HCB
236 HCB
345 HCB
Clophen A-60
Toxic Effect Observed
Reproductive
Reproductive
Reproductive
Reproductive
Reproductive/Kit survival
Reproductive
Kit growth
Reproductive
Kit growth
Reproductive
Reproductive
Reference
Linzey, 1988
Bleavins et al.
1980
Bleavins et al.
1980
Platonow and
Karstad, 1973
Hornshaw et al.
1983
Jensen et al.
1977
Wren et al.
1987
Aulerich and
Ringer, 1977
U.S. EPA, 1980
Aulerich et al.,
1985
den Boer, 1984
                                                                                     4-5

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         In calculating a wildlife value for both mink and river otter, a species sensitivity factor
    (SSF) of 1 was used because numerous studies (Ringer et al., 1972; Platonow and Karstad, 1973;
    Jensen et al. 1977; Aulerich and Ringer, 1977; U.S. EPA,  1980; Bleavins et al. 1980) have
    demonstrated that mink are among the most sensitive mammalian species to the toxic effects of
    PCBs.

    Input parameters for the wildlife equation are presented below.
                     NOAEL (mammalian)
                   BAF3 (Trophic Level 3)
                   BAF4 (Trophic Level 4)
                                    SSF
              3.0 W/V.Q body weight/day
              1,000,000  //kg body weight
              2,800,000  t/kg body weight
              1  (mink and otter)
    Body weights (WtJ, ingestion rates (FJ, and drinking rates (WJ for mink and river otter are
    presented in Table D-2 of the methodology document (Appendix D to 40 CFR 132) and shown
    below.
                                 WtA (mink)
                                 WtA (otter)
                                   FA (mink)
                                   FA (otter)
                                  WA (mink)
                                  WA (otter)
                 1.0kg
                 8.0kg
                 0.15 kg/day
                 0.9 kg/day
                 0.099 I/day
                 0.64 t /day
    The equations and calculations of mammalian wildlife values are presented below.
     WV (mink)


     WV (mink)

     WV (mink)
           [NOAEL x SSF] x WtA(n>ink,
          WA(mHO + [(1.0){FA(nirtt)xBAF3)]
             (3.0//g/kg/d x 1) 1.0kg
  0.099 lid + [(1.OHO. 15 kg/d x 1,000,000 //kg)]
                    20 pg//
     WV (otter)
                                              [NOAEL x SSF] x WtA/omr)
WA(on., + l(0.5)(FA(OMr| x BAF3) + (0.5)(FA(OW) x BAF4)]
                                               (3.0//g/kg/d x 1) 8.0kg
     WV (otter)   = o.64 //d + [(0.5) (0.90 kg/d x 1,000,000 I/kg) + [(0.5) (0.90 kg/d x 2,800,000 //kg)
     WV (otter)
                    14 pg/l
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    The geometric mean of these two mammalian wildlife values results in
                             WV (mammalian)   =   e* WV("ilk) + h wv(Ber>'/2)

                             WV (mammalian)   =   e* M "" +

                             WV (mammalian)   =   17 pg/l.
Ill
 iv.  Sensitivity Analysis for Mammalian  Wildlife Value
      The values of the various parameters used to derive the mammalian wildlife value presented
 above represent the most reasonable assumptions. The purpose of this section is to illustrate the
 significance of these assumptions and the variability in the mammalian wildlife value if other
 assumptions are made for the values of the various parameters from which the mammalian
 wildlife value is derived. The intent of this section to let the risk manager know, as much as
 possible, the influence on the magnitude of the mammalian wildlife value of the assumptions
 made in its derivation.
      In deriving the PCB mammalian wildlife value, it was assumed that 100 percent of the mink
 diet was comprised of fish, although this may not necessarily be the case. This assumption may
 lead to an overestimate of PCB exposure for mink that are not primarily foraging for fish and
 aquatic invertebrates. As indicated in the Technical Support Document for Wildlife Criteria
 (Appendix to the Preamble to 40 CFR 132), the fish content of a mink diet can vary from less
 than 50 percent to the 100 percent assumed in the mink wildlife value derivation presented above.
 If it were assumed only 50 percent of a mink's diet was from aquatic resources and the remaining
 50 percent of the diet was uncontaminated, the estimated PCB exposure would be reduced by a
 factor of 2. The resulting wildlife value for the mink would be 40 pg/l, and the mammalian
 wildlife value would be 24 pg/l, rather than the mammalian wildlife value of 17 pg/l.


,  Calculation  of Avian Wildlife  Value
 /. Acute Toxic/ty
      Birds have been shown to be more resistant than mammalian species to the acute toxic
 effects of PCBs. PCB doses greater than 200 ppm in the diet (10 mg/kg body weight) caused
 some mortality among northern bobwhite (Colinus Virginians), mallards (Anas platyrhynchos), and
 ring-necked pheasants (Phasianus colchicus). PCBs provided to these birds at dietary
 concentrations of 1500 ppm (100 mg/kg body weight) have caused extensive mortality (Eisler,
 1986). Exposure to PCBs has caused  some mortality among all the avian species tested, with
 lethal concentrations depending on the length of exposure and the particular PCB mixture
 (Aulerich et al., 1973). Values of LDy, for various avian species provided with dietary
 concentrations of PCBs have varied from 604 mg/kg for the northern bobwhite to more than 6000
 mg/kg for the Japanese quail (Coturnix japonica) (Heath et al., 1972), while mallards had LCX
 values of more than 2000 mg/kg (NAS, 1979). Acute toxicity values for avian species are
 summarized in Table 4-3.

                         Table 4-3. Avian Acute Toxicity Values
PCB Congener
1221
1242
Species
Northern bobwhite (Colinus virgin/anus)
Ring-necked pheasant (Phasianus colchicus)
Japanese quail (Coturnix japonica)
Northern bobwhite (Colinus virginianus)
Mallard (Anas platyrhynchos)
LDBO (mg/kg)
> 6000
> 4000
> 6000
2098
> 2000-3182
                                                                                         4-7

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                         Table 4-3. Avian Acute Toxicfty Values (Cont.)
PCB Congener
1242 (Cont.)
1254
1260
Species
Ring-necked pheasant (Phasianus colchicus)
Japanese quail (Coturnix japonica)
Northern bobwhite (Colinus virginianus)
Mallard (Anas platyrhynchos)
Ring-necked pheasant (Phasianus colchicus)
Japanese quail (Coturnix japonica)
European starling (Sturnus vulgaris)
Red-winged blackbird (Agelaius phoeniceus)
Brown-headed cowbird (Molothrus ater)
Northern bobwhite (Colinus virginianus)
Mallard (Anas platyrhynchos}
Ring-necked pheasant (Phasianus colchicus)
Japanese quail (Coturnix japonica)
LDM (mg/kg)
2078
> 6000
604
> 2000 - 2699
1091
2898
1500
1500
1500
747
1 975 - > 2000
1260
2186
 Sourc*: Eieler (1986).

         For all avian species, PCB residue concentrations of at least 310 mg/kg fresh weight in the
    brain were associated with an increased likelihood of death from PCB poisoning (Eisler, 1986).
    Residues in brains of starlings (Sturnus vulgaris), red-winged blackbirds (Agelaius phoeniceus),
    common grackles (Quiscalus quisculd), and brown-headed cowbirds (Molothrus ater) that died
    after ingesting diets containing 1500 ppm of Aroclor 1254 ranged from  349 to 763 mg/kg. Brains
    of birds surviving at the 50 percent mortality point contained 54 to 301  ppm PCBs (Stickel et al.
    1984).

    //'. Chronic  Toxic/ty
         Chronic toxicity studies have been conducted on mallards, Japanese quail, pheasants, and
    domestic leghorn chickens (Callus). Chickens have been shown to be more sensitive to the effects
    of chronic exposure to PCBs than have other avian species.
         Custer and Heinz (1980) fed 9-month-old mallards with a dietary dosage of 25 ppm Aroclor
    1254 for at least a month before egg-laying. Treatment did not affect reproductive success or nest
    attentiveness during incubation. The number of hens laying,  date of the first egg laid, clutch size,
    survival of ducklings to 3 weeks of age, and the number of times off the nest per day and total
    time off the nest per day did not differ between the exposed  group and the controls. Fertility of
    eggs was greater among the treated birds than among controls, a phenomenon that the authors
    attributed to males coming into reproductive condition sooner as a result of the PCBs. Using a
    mallard body weight of 1 kg (Delnicki and Reinecke,  1986)  and food ingestion rate of 0.058
    kg/day derived from the allometric equation provided in the  methodology document (Appendix D
    to 40 CFR 132), the treatment concentration can be calculated to be equivalent to a dose of 1.45
    mg/kg/day.
         In contrast to the results of the mallard study, dietary exposure to  PCBs had marked  effects
    among chickens at the same or lower concentrations. Britton and Huston (1973) exposed white
    leghorn hens to Aroclor 1242 at 0, 5, 10, 20, 40, and 80 ppm in a commercial feed  over a 6-
    week period. Following treatment,  the hens were held for an additional 6 weeks on a PCB-free
4-8

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diet and effects on reproduction were assessed. Dietary PCBs did not alter egg weight, shell
thickness, or shell weight over the 12-week experiment. PCBs in the diet did have an effect on
the hatchability of eggs. By the second week, no eggs laid by hens fed 80 ppm PCBs hatched.
Hatchability unproved as the concentration of PCBs in the diet decreased. A significant reduction
in hatchability of the eggs laid by hens fed 10 ppm Aroclor 1242 was observed at the sixth week
of the experiment, but no effect on hatchability was noted for the eggs  laid by hens fed a 5 ppm
diet. Using a chicken weight of 1.66 kg (Lillie et al. 1975; and personal communication with Dr.
Wayne Kunzel, Poultry Science Department, University of Maryland)  and a food ingestion rate of
0.81 kg/day derived from the allometric relationship presented in the methodology document
(Appendix D to 40 CFR  132), the NOAEL for Aroclor 1242 determined  from this study was
calculated to be 2.44 mg/kg/day (5 ppm) for hatchability of eggs.
     Aroclor 1254 was also found to cause reduced egg production and hatchability in chickens.
In a subchronic study, Platonow and Reinhart (1973) fed chickens rations containing 0, 5, or 50
ppm Aroclor 1254 for up to 39 weeks. A drastic decline in production  and hatchability of fertile
eggs was observed among hens maintained at the 50 ppm level. At 5 ppm, egg production was
reduced, but not the hatchability of the fertile eggs. Fertility for the 5 ppm group was similar to
the control during the first 14 weeks, but declined significantly in the last 14 weeks. These results
indicate a LOAEL of 5 ppm for egg production  and fertility. Using the chicken body weight and
feed ingestion rate presented above, the LOAEL was calculated to be 2.44 mg/kg/day.
     Lillie et al. (1975) assessed the reproductive effects of various PCBs (i.e., Aroclors 1232,
1242, 1248, 1254, and 1016) on white leghorn chickens maintained on a  commercial feed treated
at 0, 2, 5, 10, and 20 ppm total PCBs for 8 to 9 weeks. The data presented by Lillie et al. (1975)
were pooled, both across Aroclors and across dose rates, making their interpretation unreliable.
However, the data indicate no effect on egg production from dietary exposure at any
concentration of any of the Aroclors. Furthermore, the data indicate that PCB levels of 5 ppm hi
feed,  regardless of congener, has no effect on hatchability, while Aroclors 1232, 1242 and 1248,
regardless of concentration, but probably at 10 and 20 ppm, caused reduced hatchability. None of
the Aroclors or dose levels had any  effect on egg weight, eggshell thickness, adult body weight
changes, feed consumption, livability, or fertility.
     In another paper Lillie et al.  (1974) found that dietary exposure to either 2 or 20 ppm of the
various PCBs had no effect on adult body weight, adult mortality, fertility, egg weight, or
eggshell thickness. Reduced egg production and  egg hatchability were observed among the
different groups of chickens maintained on 20 ppm Aroclor 1232, 1242, 1248, or 1254. These
effects were not observed at a dietary concentration of 2 ppm. Lillie et al. (1974) also monitored
the growth and survival of chicks  produced from hens maintained on Aroclor-treated feed. A
significant reduction in growth was observed among chicks produced from hens maintained on
feed treated with either Aroclor 1248 or Aroclor 1254 at 2.0 and 20 ppm. Significant reduction in
weight gain was  also observed among  chicks produced from hens maintained on feed treated with
either Aroclor 1232 or Aroclor 1242 at 20 ppm but not at 2 ppm. Only Aroclor 1248 at a
concentration of 20 ppm in the maternal diet was associated with significant chick mortality. The
results of this study indicate a 2.0 ppm NOAEL and a 20 ppm LOAEL for egg production and
hatchability with Aroclors 1232, 1242, 1248, or 1254.  In addition, a 2.0 ppm LOAEL for chick
growth effects for Aroclor 1248 and 1254, and a 2.0 ppm NOAEL for  Aroclors 1232 and 1242
can be inferred. Using the chicken body weight and food ingestion rates presented previously, the
LOAEL and NOAEL for egg production and hatchability can be calculated to be 9.8 and 0.98
mg/kg/day, respectively.  For chick growth effects, the LOAEL for Aroclors 1248 and 1254, and
the NOAEL for Aroclors 1232 and 1242 are 0.98 mg/kg/day.
     Scott (1977) measured the effect of Aroclor 1248 on reproductive parameters of white
leghorn hens maintained at dietary concentrations of 0.5, 1.0, 10.0, and 20.0 ppm over an 8-
week period.  A significant reduction in egg production at the 20 ppm concentration and a
decrease in hatchability of fertile eggs  at the  10 ppm dose after  8 weeks were noted. No
significant effects on these reproductive endpoints were observed at 1 ppm Aroclor 1248 in the
                                                                                        4-9

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    diet. Using the chicken body weight and food ingestion rate presented above, the NOAEL for
    reproduction is 0.49 mg/kg/day, and the LOAEL is 4.9 mg/kg/day.
         Dahlgren et al. (1972) assessed the effects of orally-administered Aroclor 1254 on
    reproduction in the ring-necked pheasant. Female pheasants were individually dosed once per
    week, via gelatin capsule, at rates of 0, 12.5, and 50 mg/week; and male pheasants were dosed at
    rates of 0 and 25 mg/week, for 16 weeks. Egg production, egg fertility, egg hatchability,
    survivability, and growth of chicks through 6 weeks post-hatch were monitored.  Egg production
    and chick survivability were significantly reduced among hens  administered 50 mg Aroclor 1254
    per week, but not among hens administered 12.5 mg per week. No effect of Aroclor 1254
    administration on egg fertility was noted, although significant reductions in hatchability were
    reported among eggs from both treatment groups. No effect of Aroclor treatment on chick growth
    was observed.  Using a pheasant body weight of 1 kg (John Nosek, personal communication), a
    value of 1.8 mg/kg/day can be inferred from this study for the NOAEL for egg production and
    chick survivability, and for the LOAEL for egg hatchability.
         The various toxicity values derived from the studies discussed above are summarized in
    Table 4-4. An evaluation of these studies suggest mat the lowest LOAEL values  are those for
    chick growth from chickens dosed with Aroclors 1248 and 1254 (Lillie et al.  1974) and the value
    for egg hatchability among pheasants obtained with Aroclor 1254 (Dahlgren et al. 1972).  The
    lowest NOAELs  were for egg production and hatchability among chickens using Aroclors 1232,
    1242, 1248, or 1254 (Lillie et al. 1974; Scott, 1977).
         The results  of the pheasant study by Dahlgren et al. (1972) are used to derive the avian
    wildlife value.  According to the methodology document, preference is given to laboratory studies
    with wildlife species. Pheasants have been show to be as sensitive to PCBs as chickens, the more
    traditional avian  laboratory species. The toxic endpoint of egg hatchability is a meaningful
    reproductive effect that is associated with avian dietary exposure to PCBs. In addition, the study
    by Dahlgren et al. (1972) involved exposures to both male and female adults. Calculation of the
    avian wildlife values for PCBs is based on the study of Dahlgren et al. (1972), where a LOAEL
    of 1.8 mg/kg/day for egg hatchability was determined for Aroclor 1254.

                       Table 4-4. Summary of Chronic Avian PCB Studies
Species
Mallard
Chicken
Chicken
Chicken
Chicken
LOAEL
(mg/kg/day)


2.44
4.88
4.88
4.88


9.8
9.8
NOAEL
(mg/kg/day)
1.45
2.44

2.44
2.44
2.44
2.44
2.44
0.98
0.98
PCB
Congener
Aroclor 1 254
Aroclor 1242
Aroclor 1 254
Aroclor 1 232
Aroclor 1 242
Aroclor 1 248
Aroclor 1 254
Aroclor 1016
Aroclor 1232
Aroclor 1 242
Toxic Effect
Observed
Reproduction
Egg hatchability
Egg production
and Fertility
Egg hatchability
Egg production
and Hatchability
Reference
Custer and
Heinz, 1980
Britton and
Huston, 1973
Platonow and
Reinhart, 1973
Lillie et al.
1975
Lillie et al.
1974
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                Table 4-4. Summary of Chronic Avian PCB Studies (Cont.)
Species
Chicken
(Cont.)
Chicken
Pheasant
LOAEL
(mg/kg/day)
9.8
9.8


0.98
0.98
4.9
1.8
NOAEL
(mg/kg/day)
0.98
0.98
0.98
0.98


0.49

PCB
Congener
Aroclor 1248
Aroclor 1 254
Aroclor 1232
Aroclor 1242
Aroclor 1248
Aroclor 1 254
Aroclor 1 248
Aroclor 1 254
Toxic Effect
Observed
Egg production
and Hatchability
Chick growth
Egg production
and Hatchability
Egg hatchability
Reference
Lillie et al.
1974
Scott, 1977
Dahlgren et al.
1972
///. Avian  Wildlife Value Calculation
     Dividing the LOAEL for egg hatchability by a LOAEL to NOAEL uncertainty factor of 10
gives a NOAEL for calculating avian wildlife values of 0.18 mg/kg/day.
     Results of the chicken and pheasant studies suggest that these 2 species are similarly
sensitive to the toxic effects of PCBs, which suggests that a 0.1 SSF may be unduly conservative
in deriving avian-specific wildlife values. In that piscivorous species may be more sensitive to
PCB toxicity than the chicken or pheasant, a SSF of 0.3, intermediate to 0.1 and 1.0, was
selected.
The wildlife equation and input parameters are presented below.
                    NOAEL (avian)
             BAF3 (Trophic Level 3)
             BAF4 (Trophic Level 4)
                              SSF
0.18 mg/kg body weight/day
1,000,000 f/kg body weight
2,800,000 I /kg body weight
0.3 (kingfisher, osprey and eagle)
Values for body weights (WtJ, ingestion rate (F^, and drinking rate (WJ for kingfisher, osprey
and eagle are presented in Table D-2 of the methodology document (Appendix D to 40 CFR
132), and shown below.
                        WtA (kingfisher)
                           WtA (osprey)
                             WtA (eagle)
                          FA (kingfisher)
                             FA (osprey)
                              FA (eagle)
                         WA (kingfisher)
                            WA (osprey)
                             WA (eagle)
     0.15 kg
     1.5kg
     4.5kg
     0.075 kg/day
     0.3 kg/day
     0.5 kg/day
     0.017 f/day
     0.077 f/day
     0.16 f/day
                                                                                    4-11

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    Calculations of avian wildlife values are summarized below.

                                           (NOAEL x SSF) x WtA(tinB(irf,,
             WV (kingfisher)



             WV (kingfisher)


             WV (kingfisher)



                WV (osprey)    =



                WV (osprey)    =


                WV (osprey)    =



                 WV (eagle)
                                            x BAF3)]

             	(O.ISmg/kg/d x 0.3) 0.15 kg	
             0.017 tld + [{1.0X0.075 kg/d x 1.000,000 //kg)]

                               110pg//


                        (NOAEL x SSF) x WtA(owyl
                              [d.O)(FA(owl x BAF3)]
                       (0.18 mg/kg/d x 0.3) 1.5kg
              0.077 */d + [d.0)(0.3 kg/d x 1,000,000 //kg)]
                               270 pglt
                        (NOAEL x SSF) x WtAlMBl>,
                               [(1.0)(FAMOl.xBAF4)

                       (0.1 8 mg/kg/d x 0.3) 4.5 kg
                                                                      _
                 WV (eagle)    =       0.16 //d + [(1.0K0.5 kg/d x 2,800,000 //kg)]

                 WV (eagle)    =                       170pg//

    The geometric mean of these three avian wildlife values results in:
WV

WV

WV (avian)
                                                     + In WV(o*pry) + In WVIgMI/3l


                                                  + ln270pg/J + ln170pg/ll/3)
                                         170pg/l.
    iv.  Sensitivity Analysis of Avian Wildlife Value
         The values of the various parameters used to derive the avian wildlife value presented above
    represent the most reasonable assumptions. The purpose of this section is to illustrate the
    significance of these assumptions and the variability in the avian wildlife value if other
    assumptions are made for the values of the various parameters from which the avian wildlife
    value is derived. The intent of this section is to let the risk manager know,  as much as possible,
    the influence on the magnitude of the avian wildlife value of the assumptions made in its
    derivation.
         No chronic PCB toxicity studies using piscivorous avian species were identified; however, it
    could be assumed that such species are more sensitive to the effects of PCBs than the 0.3 SSF
    would suggest. Use of a SSF of 0.1 would result in an avian wildlife value of 57 pg/l instead of
    170 pg/f.
         Chickens have been shown to be among the most sensitive species to PCB toxicity. Chronic
    toxicity studies with chickens suggest effects on reproductive success could be expected at a
    threshold between 0.24 and 0.98 mg/kg/day (Lillie et al. 1974, 1975).  Using these values as the
    NOAEL hi calculating avian  wildlife values, and using a SSF of 0.3 yields avian wildlife values
    of 230 pg/f to 940 pg/f, respectively. The use of a SSF of 0.1 would result in avian wildlife
    values ranging from 76 pg/f  to 310 pg/f instead of the avian wildlife value of 170 pg/f.
4-12

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         Mallard studies are also available to calculate wildlife values, and these may be considered
    more representative of sensitive wildlife species than those from chicken or pheasant.  Mallard
    studies yield a NOAEL for reproduction of 1.45 mg/kg/day (Custer and Heinz, 1980). If the
    results of the mallard study were used with a SSF of 0.3, the avian wildlife value would be
    approximately 1400 pg/l. If the mallard NOAEL were used with a SSF of 0.1, the avian wildlife
    value would be approximately 460 pg/l, instead of the avian wildlife value of 170 pg/l.
         The derivation of an avian wildlife value is based on the assumption that 100 percent of an
    eagle's diet is composed of fish. A study by Kozie and Anderson (1991) suggests that fish
    comprise 97 percent of Lake Superior uigle diets, and mammals and birds each comprise l.S
    percent of eagle diets. Assuming the metabolizable energy in fish is approximately 1 kcal/g
    (Palmer,  1988; and  Stalmaster and Gessaman, 1982) and the typical eagle consumes about 500 g
    of fish per day (Technical Support Document for Wildlife Criteria, Appendix to the Preamble to
    40 CFR 132), an eagle has a daily energy requirement of 500 kcal/day. The energy content for
    birds is 2 kcal/g (a value derived for mallards; Stalmaster and Gessaman, 1982). Applying the
    conservative assumptions that the bioaccumulation in mammals would be equivalent to that in
    Trophic Level 4 fish and the caloric value would be the  same for mammals and fish, an eagle diet
    consisting of 1.5 percent fish-eating birds and 98.5 percent fish would result in a daily intake of
    approximately 7.4 g of bird and 480 g of fish to meet the daily energy requirement of 500
    kcal/day. Braune and Nordstrom (1989) have reported that total PCBs bioaccummulate in Lake
    Ontario herring gulls at a level approximately 90 times higher than that observed in alewife.
    Therefore, dietary exposure to eagles of total PCBs would be higher if piscivorous birds comprise
    a portion of their diets. The total PCBs exposure to eagles eating 7.4 g of piscivorous birds a day
    would be approximately 2.3 times higher man an exposure associated with a 100 percent fish diet.
    Such an analysis would result in a bald eagle wildlife value of 75 pg/l, and an avian wildlife
    value of 130 pg/l  compared to 170 pg/l.


IV. Great Lakes  Wildlife Criterion
         The Great Lake Wildlife Criterion for polychlorinated biphenyls (PCBs) is determined by
    the lower of the mammalian wildlife value  (17 pg/l) and the avian wildlife value (170 pg/l). The
    mammalian wildlife value is one order of magnitude  smaller than the avian wildlife value.
    Therefore, the Great Lake Wildlife Criterion for polychlorinated biphenyls (PCBs) is 17 pg/l.


    /. Discussion of Uncertainties
         Wildlife populations inhabiting the Great Lakes basin would not be impacted from the intake
    of drinking water and aquatic prey taken from surface water containing PCBs in concentrations of
    17 pg/l, based on the uncertainty factors used to account for data gaps and the variability in the
    toxicity and exposure parameters inherent in the PCB risk assessment.  Criteria for other
    ecoregions may require an  analysis of different wildlife species with different diets and body
    masses. In addition, the bioaccumulation factors in this analysis were based on an analysis for the
    Great Lakes,  and different  bioaccumulation factors may be more appropriate for other
    waterbodies.
         Finally, generic assumptions were made in assessing the hazards of PCBs to wildlife
    populations through  the use of LOAELs and NOAELs for reproduction and development. The use
    of these levels assumes no hazards to  wildlife populations would result from the direct exposure
    of individuals to PCBs. However, it could be argued that some increase in density independent
    mortality, or decrease in density independent reproductive success,  which could be attributable to
    exposure to PCBs, could be incurred without impacting the population dynamics of a species. In
    general, well-validated population models do not yet exist for the species analyzed, and it is
    difficult to estimate the extent of mortality or reproductive failure that could be incurred. In
    addition, the interaction of additional chemical as well as non-chemical stressors on wildlife
    population responses is also poorly resolved at this time.
                                                                                          4-13

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     mink. Arch. Environ. Con tarn. Toxicol.  15:393-399.
 Aulerich, RJ., S. J. Bursian, W. J. Breslin, B. A. Olson, R. K. Ringer. 1985. lexicological manifestations of
     2,4,5,2',4',5'-; 2,3,6,2',3',6'-; 3,4,5,3',4',5'- hexachlorobiphenyl and Aroclor 1254 in mink. J. Toxicol. Environ.
     Health 15:63-79.
 Aulerich, R. J. and  R. K. Ringer. 1977. Current status of PCB toxicity to mink, and effect on their reproduction.
     Arch. Environ. Contain. Toxicol. 6:279-292.
 Aulerich, RJ., R.K. Ringer and S. Iwamoto. 1973. Reproductive failure and mortality in mink fed on great lakes fish.
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 Bleavins, M. R., R. I. Aulerich and R. K. Ringer. 1980. Polychlorinated biphenyls (Aroclor 1016 and 1242): Effects
     on survival and  reproduction in mink and ferrets. Arch.  Environ. Contain. Toxicol.  9:627-635.
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     Proceedings 36(6): 1888-1893.
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Stalnuster, M. V., and I. A. Gessaman.  1982. Food consumption and energy requirements of captive bald eagles. J.
    Wildl. Manage. 46:646-654.
Stickd, W.H., L.F. Stickel, R.A. Dyrland and D.L. Hughes. 1984. Aroclor 1254* residues in birds: Lethal levels and
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