Appendix G
Toxicity of PCBs
This appendix summarizes available data on PCB toxicity that are potentially useful for a risk assessment; it is not meant to provide a comprehensive review of PCB toxicity data. Data related to humans and aquatic organisms, such as freshwater and marine invertebrates, fish, birds, and marine mammals, that could be exposed to PCB-contaminated sediments are evaluated. The limitations of these toxicity data are also discussed, followed by a discussion of how this information is used in risk assessments, including toxic equivalency factors.
TOXIC EFFECTS OF PCBs
The toxicity of PCBs is well established from laboratory and field studies (Giesy et al. 1994a,b). Chronic toxicity has been observed in fish, birds, and mammals. Several studies demonstrate, however, that individual PCB congeners can act through different mechanisms and have different toxic potentials (Safe 1984, 1994; Strang et al. 1984; Seegal 1996). The overall impacts of PCBs on the environment and biota are due to not only the individual components of the mixtures but also the interactions (additive, synergistic, and antagonistic) among the PCB congeners present and between the PCBs and the other chemicals. Risk assessments of PCBs, therefore, require information on the levels of individual PCB congeners present in the PCB mixture and data on their interactions. Developments in high-resolution isomer-specific PCB analysis have made identification and quantitation of individual PCB conge-
ners feasible, but challenges remain in the risk-assessment process because of the differing toxicity of individual PCBs.
Most in vivo animal studies and in vitro bioassays use commercially available, technical-grade mixtures of PCBs or individual congeners. PCBs present in the environment differ from commercially available mixtures, because different congeners are metabolized and biodegraded at different rates. Few studies have investigated the effects of environmentally altered mixtures of PCBs. The studies that have investigated those effects include field and controlled laboratory feeding experiments, but the co-occurrence of other toxicants, such as DDT, Toxaphene, and dieldrin, complicate their interpretation.
Commercial PCB mixtures elicit a broad spectrum of toxic responses that depend on several factors, including chlorine content, purity, dose, species, strain, age, and sex of animal, and route and duration of exposure. Immunotoxicity, carcinogenicity, neurotoxicity, and developmental toxicity, as well as the biochemical effects of commercial PCB mixtures, have been extensively investigated in various laboratory animals, fish, and wildlife species. The mechanisms and endpoints of PCB toxicity have been reviewed (Poland and Knutson 1982; Safe 1984; Barrett 1995; Silberhorn et al. 1990). Two main categories of PCBs have been designated based on mechanism of action: those that act through the arylhydrocarbon receptor (AhR) and those that do not.
AhR-Mediated Effects
The non- and mono-ortho-substituted PCBs are of particular concern, because these congeners can assume a planar or nearly planar conformation similar to that of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (Safe 1990; Giesy et al. 1994a; Metcalfe and Haffner 1995) and have toxic effects qualitatively similar to TCDD. These compounds act by the same mechanism as TCDD, that is, by binding to and activating the AhR, a cytosolic, ligand-activated transcription factor (Poland and Knutson 1982; Gasiewicz 1997; Blankenship et al. 2000). Each polychlorinated dibenzo hydrocarbon (PCDH) binds with different affinity to the AhR and, therefore, has different potency for biological effects (Safe 1990; Ahlborg et al. 1994; van den Berg et al. 1998).
A great deal of work has been conducted on the toxicity of dioxin, and there is a large amount of data on dioxin-like effects. The data come from experimental and epidemiological studies. Epidemiological studies have been conducted in individuals exposed to dioxin occupationally, in the chemical and the agricultural industries, individuals exposed environmentally (e.g.,
following an accident in Seveso, Italy, people living in farming communities were exposed to high levels of dioxin), and individuals exposed to Agent Orange, a herbicide that contained dioxin as a contaminant, during the conflict in Vietnam. Dioxin-like effects have been reviewed in great detail in other reports. In addition to numerous published review articles, a review was published by the Institute of Medicine (IOM) on the toxicity and epidemiological data on AhR-mediated effects in Veterans and Agent Orange: Health Effects of Herbicides Used in Vietnam and its subsequent updates (IOM 1994, 1996, 1999, 2000, 2001). EPA has also issued a draft health risk assessment of TCDD that summarizes and reviews the literature on the toxicity and epidemiology of dioxin and dioxin-like compounds (EPA 2000). Therefore, AhR-mediated health effects are only mentioned briefly here, and the reader is referred to other sources for a summary of the numerous studies.
As summarized by IOM (2001), dioxin-like effects comprise a diverse spectrum of sex-, strain-, age-, and species-specific effects, including carcinogenicity, immunotoxicity, reproductive or developmental toxicity, hepatotoxicity, neurotoxicity, chloracne, and loss of body weight. The wasting syndrome occurs at high concentrations of TCDD and is characterized by a loss of body weight and fatty tissue. Dioxin causes toxicity in the liver where lethal doses of TCDD cause necrosis (i.e., cell death). Effects on the morphology and the function of the liver are seen at lower doses. Dioxin affects the endocrine system of animals. Some experiments indicate that thyroid hormone levels are altered by activation of the AhR. Some neurodevelopmental effects have been seen in rats and monkeys after in utero TCDD exposure; some of those effects are not mediated by the AhR. In animals, one of the most sensitive systems to AhR-mediated toxicity is the immune system. Recent studies have demonstrated that dioxin can alter the levels of immune cells, the measured activity of those cells, and the ability of animals to fight off infection. Effects on the immune system, however, appear to be dependent on the species and strain of animal studied. Reproductive and developmental effects have been seen in animals exposed to dioxin. Effects on sperm counts, sperm production, and seminal vesicle weights have been seen in male offspring of rats treated with TCDD during pregnancy. Effects have also been seen on the female reproductive system following developmental exposure to TCDD. In some recent studies, however, the effects on the male and female reproductive systems were not accompanied by effects on reproductive outcomes. TCDD is a potent tumor promoter in laboratory rats. Evidence for an association between dioxin and some cancers (soft-tissue sarcoma, non-Hodgkin’s lymphoma, Hodgkin’s disease, and respiratory disease) is seen in epidemiological studies. Recent epidemiological studies also suggest an association between TCDD exposure and an increased incidence of diabetes.
Non-AhR-Mediated Effects
PCBs with two or more ortho chlorines do not interact with the AhR and elicit a different pattern of toxicity. These PCBs have been shown to elicit a diverse spectrum of non-Ah-receptor-mediated toxic responses in experimental animals, including neurobehavioral (Bowman et al. 1978, 1981; Schantz et al. 1991), neurotoxic (Fingerman and Russel 1980; Seegal et al. 1990; Seegal 1996; Kodavanti and Tilson 1997), carcinogenic (Barrett 1995; Ahlborg et al. 1995), and endocrine changes (Brouwer 1991; Van Birgelen et al. 1995). In addition, some of the metabolites of PCBs have antiestrogenic properties (Kramer et al. 1997) and cause hypothyroidism and decreased plasma vitamin A levels (Brouwer et al. 1995; Li and Hansen 1996; Brouwer 1991). These alterations in vitamin A and thyroid hormone concentrations might significantly modulate tumor promotion and developmental and adult neurobehavioral changes (Ahlborg et al. 1992). Although AhR-mediated toxicity is peliotropic, effects due to nondioxin-like PCBs might involve multiple unrelated mechanisms of action.
Developmental and cognitive dysfunctions observed in children born to mothers who consumed PCB-contaminated rice oil in Japan (Yusho) and Taiwan (Yu-Cheng) have been associated with exposure to halogenated aromatic hydrocarbons (HAHs) (Chen and Hsu 1994; Chen et al. 1994; Masuda 1985). The rice oil in both those incidents was contaminated with a mixture of HAHs, including PCBs, polychlorinated dibenzofurans (PCDFs), and PCQs (polychlorinated quarterphenyls) (Masuda 1985), and it has been difficult to determine which contaminants in the rice oil are responsible for the persistent behavioral and cognitive developmental alterations in the exposed children. Laboratory studies of ortho-substituted PCBs indicate that the developing nervous system is sensitive to PCBs and show effects similar to those seen following the accidental PCB poisonings and described in epidemiological reports (Kuratsune et al. 1972; Hsu et al. 1988; Jacobson et al. 1990; Seegal and Schantz 1994; Huisman et al. 1995; Seegal 1996). No or a poor correlation between the presence of AhR-mediated effects, such as chloracne and hyperpigmentation, and the observed cognitive dysfunctions suggest that the alterations in neurological function in the Yusho and Yu-Cheng children might be due to the nonplanar ortho-substituted PCB congeners present in many commercial mixtures of PCBs rather than to the coplanar contaminants that interact at the AhR (Yu et al. 1991; Rogan and Gladen 1992). The mechanisms of neurotoxic effects of ortho-substituted PCBs and the behavioral effects reported in epidemiological studies have been reviewed (Maier et al. 1994; Chu et al. 1996; Chishti et al. 1996; Eriksson and Fredriksson 1996a,b; Morse et al. 1996a,b; Gasiewicz 1997; Jacobson and Jacobson 1997; Kodavanti and Tilson 1997; Wong et al. 1997). The neurotoxicological ef-
fects of technical PCB mixtures in experimental studies are summarized in Table G-1. Although the mechanism of PCB-induced neurotoxicity has not been determined, various biochemical effects have been investigated.1
Aroclor 1254 (Greene and Rein 1977; Kittner et al. 1987; Seegal et al. 1989)2 reduced dopamine (DA) content in pheochromocytoma (PC 12) cells, a continuous cell line that synthesizes, stores, releases, and metabolizes DA in a manner similar to that of the mammalian central nervous system. Although the decrease in DA in one of those studies (Angus and Contreras 1996) was attributed to the cytotoxicity, it implies that the non-ortho PCBs could be lethal to cells at concentrations that are neurotoxic for certain ortho-substituted PCBs. Studies in a mouse neuroblastoma cell line (NIE-N115) also showed effects on the dopaminergic system (Seegal et al. 1991a,b). Of about 50 individual PCB congeners tested in PC 12 cells, di-ortho- through tetra-ortho-substituted congeners were the most potent at affecting DA content, whereas coplanar PCB congeners were ineffective (Shain et al. 1991). In addition, chlorination in a meta-position decreased the potency of ortho-substituted congeners, but meta-substitution had little effect on congeners with both ortho-and para-substitutions. Changes in the content of neurotransmitters, such as DA, have been seen following exposure to PCBs in mice and nonhuman primates (Seegal et al. 1985, 1986a,b, 1991a,b,c, 1994). Those results suggest that some of the neurotoxicity associated with PCBs might be due to a mechanism independent of AhR activation. Rats appear to be less sensitive to the effects of PCBs on DA content (Morse et al. 1996a,b), suggesting that the potency of ortho-substituted PCBs in reducing brain DA might be species-specific. Other studies also suggest that the non-ortho coplanar congener 3,3′,4,4′-(PCB 77) alters DA concentrations in a species-, age- and dose-dependent manner (Agarwal et al. 1981; Chishti and Seegal 1992). Similarly, effects of PCBs on learning behavior appear to be sex-specific, females being more sensitive than males (Schantz et al. 1992), and age-specific, effects being prominent after prenatal exposure but less so after adult exposure (Seegal 1996). Cholinergic function was affected by early postnatal exposure of mice to non-ortho coplanar congeners (Eriksson et al. 1991; Eriksson and Fredriksson 1998).
Some PCB congeners also affect Ca2+ homeostasis and protein kinase C (PKC) translocation in cerebellar granule cells (Kodavanti et al. 1993a,b, 1994, 1996). That activity is congener-specific; ortho-substituted PCBs, but not AhR-activated congeners, affect Ca2+ homeostasis (Kodavanti et al. 1995).
TABLE G-1 Summary of Effects of Peri- and Postnatal Exposures to PCBs on Neurotoxic Effects in Animals
PCB Congener/ Mixture |
Species, Sex, Age |
Dose and Exposure |
Effects and Effective Doses |
Reference |
In vivo studies |
||||
3,3′,4,4′-(PCB 77) |
CD-1 mice, pregnant female |
32 mg/kg of birth weight (bw), oral, prenatal exposure, 10–16 d of gestation |
Hyperactivity in offspring, neuromuscular dysfunction, learning and performance deficits, ‘spinning’ syndrome |
Tilson et al. 1979 |
3,3′,4,4′-(PCB 77) |
CD-1 mice, pregnant female |
32 mg/kg bw, oral, prenatal exposure, 10–16 d of gestation |
Hyperactivity in offspring, reduction in brain dopamine, behavioral alterations |
Agarwal et al. 1981 |
3,3′,4,4′-(PCB 77) |
NMRI mice, male, 10 d |
0.41–41 mg/kg bw, oral, single postnatal exposure |
Cholinergic system affected at 0.41 mg/kg bw, disturbed behavior |
Eriksson et al. 1991 |
2,4,4′-(PCB 28) |
NMRI mice, male, 10 d |
0.18, 0.36, 3.6 mg/kg bw, oral, single postnatal exposure |
After 4 months aberrations in spontaneous behavior, lack of effect on memory and learning and on nicotinic receptors, no effect on dopamine or serotonin, 0.36 mg/kg bw reduced total activity |
Eriksson and Fredriksson 1996b |
2,2′,5,5′-(PCB 52) |
NMRI mice, male, 10 d |
0.2, 0.41, 4.1 mg/kg bw, oral, single postnatal exposure |
After 4 months aberrations in spontaneous behavior, deficits in memory and learning function, cholinergic nicotinic receptors affected, no effect on dopamine or serotonin, 4.1 mg/kg bw reduced total activity |
Eriksson and Fredriksson 1996b |
2,3′,4,4′,5-(PCB 118) |
NMRI mice, male, 10 d |
0.23, 0.46, 4.6 mg/kg bw, oral, single postnatal exposure |
No significant changes in spontaneous and swim-maze behavior up to the dose of 4.6 mg/kg bw |
Eriksson and Fredriksson 1996b |
2,3,3′,4, 4′,5-(PCB 156) |
NMRI mice, male, 10 d |
0.25, 0.51, 5.1 mg/kg bw, oral, single postnatal exposure |
No significant change in spontaneous and swim-maze behavior up to the dose of 4.6 mg/kg bw |
Eriksson and Fredriksson 1996b |
2,2′,5,5′-(PCB 52) |
NMRI mice, male, 10 d |
4.1 mg/kg bw, oral, single postnatal exposure |
At 4 months decrease in rearing, locomotion and total activity |
Eriksson and Fredriksson 1996a |
3,3′,4,4′,5-(PCB 126) |
Sprague-Dawley rats, both sexes, 5–7 wk (weanling) |
0.1–100 ng/g in diet for 13 wk, oral, postnatal |
Growth suppression, thymic atrophy, increased liver weight, anemia, no significant alterations in biogenic amines, NOAEL=0.1 ng/g in diet or 0.01 mg/kg bw/d |
Chu et al. 1994 |
3,3′,4,4′-(PCB 77) |
Sprague-Dawley rats, both sexes, 5–7 wk (weanling) |
10–10,000 ng/g in diet for 13 wk, oral, postnatal |
Increased EROD activity, decreased vitamin A, altered dopamine and homovanillic acid in brain, histopathological changes in thyroid and liver, NOAEL=100 ng/g in diet or 8.7 mg/kg bw/d |
Chu et al. 1995 |
2,3′,4,4′,5-(PCB 118) |
Sprague-Dawley rats, both sexes, 5–7 wk (weanling) |
10–10,000 ng/g in diet for 13 wk for males, 2–2000 ng/g for females, oral, postnatal |
Increased EROD activity, reduced dopamine and homovanillic acid in brain, histopathological changes in thyroid and liver, brain residues at the highest dose 0.36–1 mg/g, NOAEL=200 ng/g in diet or 17 mg/kg bw/d |
Chu et al. 1995 |
2,2′,4,4′,5, 5′-(PCB 153) |
Sprague-Dawley rats, both sexes, 5–7 wk (weanling) |
50–50000 ng/g in diet for 13 wk, oral, postnatal |
Increased EROD activity, reduction in hepatic vitamin A, decreased dopamine and its metabolites, females more sensitive, histological changes in thyroid and liver, highest dose brain residues 16–29 μg/g, NOAEL=500 ng/g in diet or 34 μg/kg bw/d |
Chu et al. 1996 |
2,2′,3,3′,4, 4′-(PCB 128) |
Sprague-Dawley rats, both sexes, 5–7 wk (weanling) |
50–50000 ng/g in diet for 13 wk, oral, postnatal |
Increased EROD activity, reduction in hepatic vitamin A, decreased dopamine and its metabolites, females more sensitive, histological changes in thyroid and liver, highest dose brain residues 5–10 μg/g, NOAEL=500 ng/g in diet or 42 μg/kg bw/d |
Lecavalier et al. 1997 |
PCB Congener/ Mixture |
Species, Sex, Age |
Dose and Exposure |
Effects and Effective Doses |
Reference |
3,3′,4,4′,5-(PCB 126) |
Lewis rats, adult female |
10 and 20 μg/kg bw on days 9, 11, 13, 15, 17 and 19 days of gestation, oral, prenatal |
Fetotoxicity, delayed physical maturation, reduced body weight in offspring, increased liver weight and EROD activity, no effect on learning or neurobehavioral performance, no residues in brain, exhibited sex differences in neurotoxicity |
Bernhoft et al. 1994 |
3,3′,4,4′,5-(PCB 126) |
Lewis rats, adult female |
2 μg/kg bw on days 10, 12, 14, 16, 18 and 20 days of gestation, oral, prenatal |
Neurotoxic effects in offspring, no fetotoxicity, behavioral alterations, hyperactivity, impaired discrimination learning, no brain residues |
Holene et al. 1995 |
2,3′,4,4′,5-(PCB 118) |
Lewis rats, adult female |
1 and 5 mg/kg bw on days 10, 12, 14, 16, 18 and 20 days of gestation, oral, prenatal |
Neurotoxic effects in offspring, no fetotoxicity, behavioral alterations, hyperactivity, impaired discrimination learning, brain residues 6–982 ng/g |
Holene et al. 1995 |
3,3′,4,4′-(PCB 77) |
Wistar rats, adult female |
1 mg/kg bw, days 7 to 18 of gestation, subcutaneous injection, prenatal |
Behavioral effects in offspring, PCB concentrations in brain 0.15 μg/g |
Weinhand-Härer et al. 1997 |
2,2′,4,4′-(PCB 28) |
Wistar rats, adult female |
1 mg/kg bw, days 7 to 18 of gestation, subcutaneous injection, prenatal |
Behavioral effects in offspring, PCB concentrations in brain 0.61 μg/g |
Weinhand-Härer et al. 1997 |
Fenclor 42 |
Fischer rats, adult female |
5–10 mg/kg bw/d intake or 25–50 mg/kg, i.p., five injections daily, 2 wk prior to mating, prenatal |
Neurotoxicity and behavioral alterations, 40 mg/kg resulted in significant postweaning behavioral effects, LOAEL=10 mg/kg bw/d |
Pantaleoni et al. 1988 |
Aroclor 1254 |
Wistar rats, adult female |
0.2–26 μg/g in diet, preweaning, perinatal exposure |
Impaired neurological development, LOAEL=2.5 μg/g |
Overmann et al. 1987 |
Aroclors 1254 and 1260 |
Wistar rats, adult male |
500–1000 mg/kg bw, single oral exposure, postnatal |
Decrease in dopamine, norepinephrine and serotonin concentrations in specific regions in brain up to 14 d after exposure |
Seegal et al. 1986b |
Aroclor 1254 |
Wistar rats, adult male |
500–1000 mg/kg bw, oral exposure for 30 d, postnatal |
Dopamine and its metabolites decreased, PCB concentrations in brain after 30 d were 75–82 μg/g, 6 di-ortho and 3 mono-ortho congeners dominated |
Seegal et al. 1991a |
Aroclor 1254 |
Wistar rats, adult female |
5 and 25 mg/kg bw from day 10 to 16 of gestation, prenatal, oral |
Alterations in seratonin metabolism in the brains of offspring after 21 and 90 d of birth, other biogenic amines (e.g., dopamine norepinephrine) in brain were unaffected, effect was significant at dose 25 mg/kg bw |
Morse et al. 1996a |
Aroclor 1254 and 3,3′,4,4′-(PCB 77) |
Wistar rats, adult female |
5 and 25 mg/kg bw from day 10 to 16 of gestation, prenatal, oral |
Reduced plasma thyroid hormone, plasma concentrations of hydroxylated metabolite of PCB 153 was greater than the 153 in fetus, neonates and weanling rats, fetus brain thyroid residues affected, effect of OH-PCBs on brain is discussed |
Morse et al. 1996b |
Clophen A30 |
Wistar rats, adult female |
5 and 30 mg/kg bw in diet or intake of 0.4 and 2.4 mg/kg/d, from 60 d prior to mating until 21 d after birth, oral |
Behavioral effects, PCDF contamination in Clophen −2.5 mg/kg, brain concentration=60 ng/g after 420 d of exposure, PCBs 28, 52 and 101 were the prevalent ones |
Lilienthal et al. 1990 |
Aroclor 1016 |
Pig-tailed macaque (Macaca nemestrina), male, 3–5 yr |
0.8–3.2 mg/kg bw/d, for 20 wk, oral, postnatal |
Persistent reduction in brain dopamine, brain PCB concentrations 1–5 μg/g, only PCBs 28, 47 and 52 accumulated in brain, lightly chlorinated PCB mixtures are more effective than heavily chlorinated ones |
Seegal et al. 1990 |
PCB Congener/ Mixture |
Species, Sex, Age |
Dose and Exposure |
Effects and Effective Doses |
Reference |
Aroclor 1260 |
Pig-tailed macaque (Macaca nemestrina), male, 3–5 yr |
0.8–3.2 mg/kg bw/d, for 20 wk, oral, postnatal |
Persistent reduction in brain dopamine, brain PCB concentrations 18–28 μg/g, di-ortho substituted hexa- and heptaCBs accumulated in brain, less effective to reduce dopamine as compared to Arolcor 1016 exposure |
Seegal et al. 1990 |
Aroclor 1248 |
Rhesus monkeys, adult female |
0.5–2.5 mg/kg in diet, exposed before and during gestation, oral, perinatal, cumulative PCB intake was 293 mg |
Hyperactivity in offspring, behavioral deficits, PCB concentrations in body fat was 20 μg/g |
Bowman et al. 1981 |
In vitro or ex vivo studies |
||||
Aroclors 1254:1260 (1:1) |
Wistar rats, male, 65 d |
10–100 μg/g in media, ex vivo brain tissue, 6 h exposure |
Decrease in dopamine and its metabolites at 20 μg/g or above, brain total PCB concentration at the effective dose was >15 μg/g |
Chishti et al. 1996 |
Aroclor 1254 |
PC-12 cells |
1–100 μg/g, in vitro, 6 h exposure |
Increase followed by a decrease in cellular catecholamine |
Seegal et al. 1989 |
2,2′-(PCB 4) |
Long-Evans hooded rats, adult male |
50–200 μM, in vitro, cerebellar granule cells exposed |
Altered Ca2+ homeostasis in cerebellar granule cells, IC50=6.17 μM, more effective than PCB 126 |
Kodavanti et al. 1993a,b |
3,3′,4,4′,5-(PCB 126) |
Long-Evans hooded rats, adult male |
50–200 μM, in vitro, cerebellar granule cells exposed |
Altered Ca2+ homeostasis in cerebellar granule cells, IC50=7.61 μM |
Kodavanti et al. 1993a,b |
2,2′-(PCB 4) |
Long-Evans hooded, male, adult rats, 40–90 d |
10–100 μM, in vitro, mitochondrial and synaptosomal preparations from brain exposed |
Mg2+-ATPase activity inhibited, but not Na+/K+-ATPase activity, ED50 is roughly 25 μM |
Maier et al. 1994 |
3,3′,4,4′,5-(PCB 126) |
Long-Evans hooded, male, adult rats, 40–90 d |
10–100 μM, in vitro, mitochondrial and synaptosomal preparations from brain exposed |
Mg2+-ATPase activity was not inhibited up to the dose of 100 μM |
Maier et al. 1994 |
2,2′,3,5′,6-(PCB 95) |
Sprague-Dawley rats, male |
1–200 μM, in vitro, microsomes of rat brain hippocampus |
Alterations in neuronal Ca2+ signal and neuroplasticity, EC50=12 mM |
Wong et al. 1997 |
2,3′,4,4′-(PCB 66) |
Sprague-Dawley rats, male |
1–200 μM, in vitro, microsomes of rat brain hippocampus |
No effect was found on [3H] ryanodine receptors suggesting no alterations in neuronal Ca2+ signal up to 200 μM |
Wong et al. 1997 |
2,2′,3,5′,6-(PCB 95), a di-ortho substituted congener, altered Ca2+ transport in rat brain microsomes (Wong et al. 1997). Another di-ortho substituted congener, 2,2′-di-CB, interfered with oxidative phosphorylation by inhibiting mitochondrial Mg2+-ATPase activity in mitochondrial and synaptosomal preparations of rat brain (Maier et al. 1994). Alterations in hormone levels involved in regulating neuronal growth and development, including thyroid hormones, could also contribute to PCB-induced neurotoxicity (Seegal 1996).
Coplanar3 HAHs inhibit estradiol-induced cell clumping in the MCF-7 breast cancer cell line (Gierthy and Crane 1984) by affecting the metabolism of estradiol to 2- and 4-hydroxy estradiol (Lloyd and Weisz 1978; Foreman and Porter 1980). Similarly, the hydroxy metabolites of PCBs are antiestrogenic (Kramer et al. 1997).
A few experimental studies have examined the effects of ortho-substituted PCBs in fish (Fingerman and Russel 1980) and birds (Kreitzer and Heinz 1974). Dietary exposure of Japanese quail (Coturnix coturnix japonica) to Aroclor 1254 at 200 μg/g for 8 d showed a suppressed avoidance response (Kreitzer and Heinz 1974). Further studies on the behavioral effects of ortho-substituted PCB congeners in fish and other wildlife are needed for risk assessment.
Few studies have examined the effects of ortho-substituted PCBs on behavioral alterations or neurotoxic effects in wildlife. Dietary exposure of mink to 2,2′,4,4′,5,5′-HxCB (PCB 153) and 2,2′,3,3′,6,6′-HxCB (PCB 136) at 5 μg/g for over 3 months did not produce significant changes in concentrations of DA, norepinephrine, or seratonin in the brain (Aulerich et al. 1985). Exposure to the di-ortho congeners 2,2′,5,5′-(PCB 52), 2,2′,4,5,5′-(PCB 101), 2,2′,3,3′,4,4′-(PCB 128), 2,2′,3,4,4′,5′-(PCB 138), 2,2′,4,4′,5,5′-(PCB 153), and 2,2′,3,4,4′,5,5′-(PCB 180) did not affect survival, growth, or reproduction in the fathead minnow Pimephales promelas, despite accumulating up to 183 μg/g (wet weight (wt)) in tissues (Suedel et al. 1997). Behavioral effects, however, were not examined in that study. Studies have demonstrated the presence of ortho-substituted tetra- through hexa-CB PCB congeners in the brains of fish. 2,2′,4,4′,5,5′-(PCB 153) was the predominant di-ortho congener; lesser chlorinated ortho-substituted congeners, such as 2,4,4′-(PCB 28) and 2,2′,5,5′-(PCB 52), did not accumulate (Qi et al. 1997). Slightly chlorinated PCBs have been shown to be metabolized in fish (Willman et al. 1997).
Studies with wildlife have demonstrated a causal link between adverse health effects and PCB exposure (Kennedy et al. 1996a,b; Giesy et al. 1994a; Bowerman et al. 1995). The observed toxicity to birds and mammals, how-
ever, correlates more strongly to TCDD equivalents (TEQs) than to total PCBs (Giesy et al. 1994a; Leonards et al. 1995).
Although neurotoxicological effects of PCBs have been seen following dietary exposure of rats, mice, or nonhuman primates with technical mixtures of PCBs, those mixtures might not represent the PCB mixtures found in environmental matrices, and the doses used in those studies are higher than those seen in the environment. Similarly, in vitro assays use high concentrations of PCBs and the EC50 values (effective concentration in 50% of the test population) for various endpoints were generally high (greater than 50 μM). The EC50 values for neurotoxicological effects in in vitro studies are presented in Table G-1.
Few studies have examined which PCB congeners are present in the brains of humans and wildlife. PCBs were not detected in brain tissues obtained from two men with Parkinson’s disease (Corrigan et al. 1996). Concentrations of total PCBs in the brain of a Yu-Cheng victim was 80 ng/g, whereas those in fat tissues ranged up to 11 μg/g (Chen and Hsu 1986). PCB 153 (2,2′,4,4′,5,5′-HxCB 1.6 ng/g, wet wt) and PCB 138 (2,2′,3,4,4′,5′-HxCB 0.96 ng/g, wet wt) were the only two congeners detected in the brain of grey seals; the concentration was only 1% of that measured in the blubber (Jenssen et al. 1996). In harbor porpoises, the PCB profile in brain tissue resembled those in other tissues, with PCB 153>PCB 138>PCB 187 (Tilbury et al. 1997), suggesting that there was no preferential enrichment of lesser chlorinated ortho-substituted PCBs in wildlife. PCB concentrations in the brains were 1.5% of those found in the blubber. Similarly, concentrations of total PCBs in the brains of marine mammals from Greek waters were 1–2% of those found in the blubber (Georgakopoulou-Gregoriadou et al. 1995). Lesser chlorinated ortho-substituted PCB congeners are metabolized in humans (Tanabe et al. 1988) and dolphins (Kannan et al. 1994; Boon et al. 1997; Leonards et al. 1997). Therefore, the accumulation of the lesser chlorinated PCBs might be small following chronic exposure. Laboratory studies have shown the presence of PCBs at greater than 1 μg/g (wet wt) in brains of exposed rats and mice, but that amount could be from exposure at higher concentrations. Therefore, the neurotoxicological effects observed in laboratory studies might occur only at higher exposures, such as those seen in Yusho and Yu-Cheng or following occupational exposures.
TOXIC EQUIVALENCY FACTORS
In the environment, PCBs usually exist as mixtures, which complicates their risk assessment. One approach to congener-specific hazard assessment for complex mixtures is to develop relative potency factors for individual conge-
ners on the basis of its mechanism of action. The complex nature of PCB mixtures found in environmental and biological samples make this a daunting, if not impossible, task. If each congener causes a different toxic response by an independent mechanism, then the relative toxicities of each congener must be determined separately. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and structurally related halogenated aromatics invoke a number of common toxic responses that are mediated through the AhR (Poland et al. 1976, 1979; Poland and Glover 1977; Poland and Knutson 1982; Safe 1990; Whitlock 1987; Goldstein and Safe 1989). TCDD has the highest affinity for the AhR and is the most toxic HAH. Structurally similar polychlorinated dibenzo-p-dioxin (PCDD), PCDF, and PCB congeners have similar effects but are less potent.
The TCDD equivalency factor (TEF) approach has been developed on the basis of AhR-binding, structure-activity relationships, and cellular responses to express the potency of various PCBs, PCDDs, and PCDFs relative to TCDD. In that way, data on TCDD, for which there is information on several endpoints in many different species, can be used to derive a maximum allowable tissue concentration (MATC) for the related comments. If congeners have the same rank order across endpoints and species and relative potencies are available for PCB congeners for a few endpoints and species, then TEFs can be developed for each congener. TCDD, the most potent AhR agonist, is designated to have a TEF of 1, and other compounds are assigned a TEF that is some fraction of 1, depending upon its characteristic responses. TEFs, however, are endpoint-and species-dependent for all PCB congeners that have been tested (Safe 1990). Regulatory agencies have established consensus TEF values for individual congeners. The World Health Organization (WHO) has proposed tentative TEFs for mammals, birds, and fish that are updated as more data become available (see Table G-2) (van den Berg et al. 1998). The mechanistic considerations for the development of TEFs for the risk assessment of PCBs, including human, teleost, and avian risk assessments, are described elsewhere (Safe 1990, 1994). Considerations for the use of mammalian, teleost, and avian TGFs are discussed briefly below, as well as the limitations of this approach.
Although TEFs have limitations and uncertainties associated with them, when used appropriately they are currently considered acceptable for use in the risk assessment of planar HAHs. Some of the uncertainties resulting from interactive effects among planar HAHs can be quantified in vitro bioassay techniques.
Mammalian TEFs
The largest amount of data for the development of TEFs comes from research in mammals. Initially, all PCB congeners were regarded as toxic,
TABLE G-2 TCDD Toxicity Equivalency Factors (TEFs) for Several Dioxin-like PCB Congeners for Fish, Birds, and Mammals
PCB Congener (IUPAC No) |
Fish TEF |
Bird TEF |
Mammal TEF |
2,3,7,8-Tetra-CDD |
1 |
1 |
1 |
3,3′,4,4′-Tetra-CB (77) |
0.0001 |
0.05 |
0.0001 |
3,4,4′,5-Tetra-CB (81) |
0.0005 |
0.1 |
0.0001 |
3,3′,4,4′,5-Penta-CB (126) |
0.005 |
0.1 |
0.1 |
3,3′,4,4′,5,5′-Hexa-CB (169) |
0.00005 |
0.001 |
0.01 |
2,3,3′,4,4′-Penta-CB (105) |
<0.000005 |
0.0001 |
0.0001 |
2,3,4,4′,5-Penta-CB (114) |
<0.000005 |
0.0001 |
0.0005 |
2,3′,4,4′,5-Penta-CB (118) |
<0.000005 |
0.00001 |
0.0001 |
2′,3,4,4′,5-Penta-CB (123) |
<0.000005 |
0.00001 |
0.0001 |
2,3,3′,4,4′5-Hexa-CB (156) |
<0.000005 |
0.0001 |
0.0005 |
2,3,3′,4,4′,5′-Hexa-CB (157) |
<0.000005 |
0.0001 |
0.0005 |
2,3,4,4′,5,5′-Hexa-CB (167) |
<0.000005 |
0.00001 |
0.00001 |
2,3,3′,4,4′,5,5′-Hepta-CB (189) |
<0.000005 |
0.00001 |
0.0001 |
Source: Data from van den Berg et al. (1998). |
and early text books suggested that the toxicity of PCB congeners was proportional to the degree of chlorination. In vivo studies conducted with rodents in the 1970s and the 1980s, however, found that the toxicity of PCB congeners varied greatly, with a small group of congeners having great toxic potential (Safe et al. 1982; Silberhorn et al. 1990). The location of chlorine atoms was found to be more important to PCB toxicity than the number of chlorine atoms. Studies using mammalian models found a correlation between the structure and the AhR-binding affinity for certain PCB congeners that exhibit “dioxin-like” activities, such as induction of cytochrome P-450s (e.g., increased arylhydrocarbon hydroxylase (AHH) activity and ethoxyresorufin O-deethylase (EROD) activity), body-weight loss, hypothyroidism, decreased hepatic or plasma vitamin A levels, porphyria, thymic atrophy, immunotoxicity, and teratogenicity (Safe 1984; Poland and Glover 1977; Goldstein and Safe 1989; Parkinson et al. 1980; Bandiera et al. 1982; Parkinson et al. 1983; Yoshimura et al. 1985; Leece et al. 1985). The non-ortho-substituted coplanar PCBs 3,4,4′,5-tetra-CB (PCB 81), 3,3′,4,4′-tera-CB (PCB 77), 3,3′,4,4′,5-penta-CB (PCB 126), and 3,3′,4,4′,5,5′-hexa-CB (PCB 169), which are substituted in both para, at least two meta, and no ortho positions,
are the most toxic PCB congeners. It is hypothesized that the lack of chlorine substitution at opposing ortho positions allows the two phenyl rings to rotate into the same plane; therefore, these congeners are commonly referred to as coplanar PCBs. The toxic potencies vary; for example, the potency ratio of PCB 126 to TCDD was 66 for body-weight loss in rats, 8.1 for thymic atrophy in rats, 10 for mouse fetal thymic lymphoid development, 125 for AHH induction in rats, and 3.3 for AHH induction in rat hepatoma H4IIE cells. Based on those data, TEFs in the range of 0.008–0.3 could be derived. A consensus mammalian TEF of 0.1 was assigned to this congener (van den Berg et al. 1998).
Similar to non-ortho coplanar PCBs, chlorobiphenyl congeners with chlorine substitution at only one ortho position (mono-ortho PCBs) can achieve partial coplanarity and exhibit AhR agonist activity. TEFs have been proposed for those PCBs based on their potency relative to TCDD in various assays (Safe 1994). Those potencies can vary by 2 to 3 orders of magnitude, depending on the species and the endpoint used to derive values. When the TEF values were calculated, in vivo studies were given more strength than biochemical changes. In mammalian studies, chronic in vivo exposures were given more weight than acute exposure studies. Current TEF values are considered tentative and subject to modification as new data become available. Recognizing the need for a more consistent approach for setting internationally accepted TEFs, the World Health Organization-European Centre for Environment and Health (WHO-ECEH) and the International Program of Chemical Safety (IPCS) initiated a project in the early 1990s to create a database containing information relevant to the setting of TEFs and to derive consensus TEFs for halogenated aromatics (Ahlborg et al. 1994). The first international TEFs for dioxin-like PCBs were proposed in 1994; those values have been revised and updated (van den Berg et al. 1998).
Teleost TEFs
Because the toxicity of coplanar PCBs vary among vertebrate taxa, recent studies have focused on determining TEFs for coplanar PCBs in fish and bird models (Bosveld and van den Berg 1994; Brunstrom et al. 1995; Newsted et al. 1995; Zabel et al. 1995; Kennedy et al. 1996a,b). The AhR is present in several fish species and cell lines (Lorenzen and Okey 1990; Swanson and Perdew 1991; Hahn et al. 1992). Therefore, the mechanistic basis for TEFs in aquatic species should be similar to that observed in mammalian systems, but at present few reports are available for the development of TEFs in fish. Fish-specific TEFs have been estimated using induction of AHH and EROD
in salmonid species (Janz and Metcalf 1991; Williams and Giesy 1992; Newsted et al. 1995), and embryo mortalities in salmonids (Walker and Peterson 1991; Zabel et al. 1995; Zabel et al. 1996) and Japanese medaka (Harris et al. 1994). Generally, TEFs for coplanar PCBs derived from early-life-stage mortality studies with rainbow trout are less than mammalian TEFs (Zabel et al. 1995). Little information is available on the toxic potencies of mono-ortho-substituted congeners in fish to speculate on appropriate teleost TEFs (Metcalf and Haffner 1995). Mono-ortho-substituted congeners (IUPAC Nos. 105, 118, and 156; 0.45 to 4.4 mg/kg) do not induce EROD activity in rainbow trout (Metcalf and Haffner 1995), and studies of cytochrome P-450 1A1 induction and mortality show that ortho-substituted PCBs lack biological activity in trout (Walker and Peterson 1991; Newsted et al. 1995; Zabel et al. 1995; Hornung et al. 1996). Congener 156, however, did induce cytochrome P-450 1A messenger RNA in a rainbow trout gonadal cell line (RTG-2), but its potency was weak relative to non-ortho-substituted congeners (Zabel et al. 1996). The use of TEFs for di-ortho PCBs derived from mammalian exposure studies, therefore, would overestimate the effects of these mixtures on fish. Fish-specific TEFs have been developed for non-ortho and mono-ortho PCBs (Walker and Peterson 1991; Walker et al. 1996) and have been tentatively adopted for use by EPA (EPA 1993). It should be noted, however, that TEF values for fish are based primarily on acute exposure data. Long-term toxicity studies with rainbow trout exposed to lower concentrations of TCDD indicate that many toxic responses can develop after a few weeks (Van der Weiden et al. 1992). Additional research is needed to refine TEFs in fish and establish consensus teleost TEFs.
Avian TEFs
Endpoints used to estimate TEFs for birds include in vitro and in ovo EROD induction (Yao et al. 1990; Bosveld et al. 1992; Kennedy et al. 1996a,b) and embryo mortalities (Brunström 1989); most are based on EROD induction potencies. Unlike rodent bioassay results (Safe 1994), PCB 169 was less potent than PCB 77 in chicken embryo hepatocytes (Brunström 1990; Kennedy et al. 1996a,b) and in embryo lethality assays in birds (Brunström 1989). TCDF was also a more potent inducer of EROD activity in avian assays than TCDD (Bosveld et al. 1997). Mono-ortho PCBs were less potent inducers of EROD than non-ortho congeners in bird models, but mono-ortho congeners were more potent in birds than in teleosts and rodents. In most environmental PCB mixtures, PCB 77 (3,3′,4,4′-tetra-CB) contributes the greatest proportion of TEQs based on avian TEFs. Therefore, more informa-
tion on the TEF and environmental fate of this congener, particularly on its pharmacokinetics in birds, is necessary for an accurate avian risk assessment.
Furthermore, avian TEFs are difficult to estimate, because there is considerable variation in the toxicity of PCB congeners among birds (Bosveld and van den Berg 1994). Although TEFs in birds are based on EROD induction with eggs and cell cultures from chickens, the preferred endpoint is embryo lethality following in ovo exposure. Domestic chickens and their embryos are considerably more sensitive to AhR-mediated responses than other avian species (Kennedy et al. 1996a,b; Lorenzen et al. 1997) and, in general, fish-eating bird species are at lest an order of magnitude less sensitive than the domestic chicken. In contrast to what was seen in mammalian and teleost cell lines, TCDF was 1.2- to 3.4-fold more potent than TCDD in the white leghorn chicken (Bosveld et al. 1997). Based on in vitro EROD induction potency of coplanar PCB congeners in several bird species, the order of sensitivity is the following: domestic chicken>ring-necked pheasant>turkey≈double-crested cormorant≈great blue heron≈ring-billed gull≈duck≈herring gull≈common tern>Foster’s tern (Sanderson et al. 1998). Therefore, toxicological information for the chicken is less appropriate than other species for risk assessment of avian wildlife species, and use of RfD values based on chicken would be over protective of most species. If chicken TEFs and RfDs are used as a surrogate for wild birds, no uncertainty factors (Ufs) should be applied.
Application of the TEF Approach
TEFs have been used to assess the risk of a mixture of PCB congeners measured in biota and or environmental matrices. The concentration of each non-ortho or mono-ortho congener detected in the biota is multiplied by its corresponding TEF to yield a TCDD-equivalent concentration, or TEQ (Table G-3) (Tanabe et al. 1989). A total TEQ for a sample can be calculated by summing the TEQs for each congener present. In that way, the toxic potential of a mixture of individual congeners can be expressed as one integrated parameter relative to the potency of TCDD.
The utility of the TEF approach to environmental risk assessment is shown by the correlation between total TEQs and adverse effects in populations of birds and fish (Tillitt et al. 1992; Giesy et al. 1994a,b). Negative correlations were reported between measured TEQs and the incidence of deformities in cormorant populations from the Great Lakes (Yamashita et al. 1993), egg volume in populations of common terns in The Netherlands (Bosveld and van den Berg 1994), survival of early life stages in populations of lake trout from the Great Lakes (Mac et al. 1993), and hatching success in Forster’s tern (Kubiak et al. 1989).
TABLE G-3 An Example for Deriving 2,3,7,8-TCDD Toxicity Equivalents (TEQs) by the TEF Approach
Congener |
TEFa |
Concentration (pg/g, wet wt)b |
TEQ (pg/g, wet wt) |
Dioxins |
|||
2,3,7,8-Tetra-CDD |
1 |
3.7 |
3.7 |
1,2,3,7,8-Penta-CDD |
1 |
6.4 |
6.4 |
1,2,3,4,7,8-Hexa-CDD |
0.1 |
3.9 |
0.39 |
1,2,3,6,7,8-Hexa-CDD |
0.1 |
34 |
3.4 |
1,2,3,4,6,7,8-Hepta-CDD |
0.01 |
33 |
0.33 |
OCDD |
0.0001 |
510 |
0.051 |
Furans |
|||
2,3,7,8-Tetra-CDF |
0.1 |
3.1 |
0.31 |
1,2,3,7,8-Penta-CDF |
0.05 |
0.5 |
0.025 |
2,3,4,7,8-Penta-CDF |
0.5 |
11 |
5.5 |
1,2,3,4,7,8-Hexa-CDF |
0.1 |
5.6 |
0.56 |
1,2,3,4,6,7,8-Hepta-CDF |
0.01 |
2.9 |
0.029 |
Non-ortho PCBs |
|||
3,3′,4,4′-Tetra-CB |
0.0001 |
350 |
0.035 |
3,3′,4,4′,5-Penta-CB |
0.1 |
330 |
33 |
3,3′,4,4′,5,5′-Hexa-CB |
0.01 |
90 |
0.9 |
Total TEQs |
|
|
54.69 |
aFrom World Health Organization (1997). bFrom Tanabe et al. (1989) for human adipose tissue. |
Limitations of TEF Approach
Interactive Effects of PCBs
Despite the ability of the TEF approach to predict the potency of some mixtures of planar HAHs, it assumes that toxic responses to planar HAHs are additive and that other classes of contaminants do not modify or add to the toxicity. Those assumptions are equivocal (Giesy et al. 1994a; Safe 1994). Some rodent data indicate that responses to mixtures of planar HAHs are
additive (Sawyer and Safe 1985; Pluess et al. 1988), but other rodent data show antagonistic (Haake et al. 1987; Biegel et al. 1989) or synergistic responses (Birnbaum et al. 1985; Bannister and Safe 1987). Recent studies have also reported additive effects (Walker et al. 1996), but other data indicated interactive effects (Pohjanvirta et al. 1995; Harper et al. 1995; Van Birgelen et al. 1995; Li and Hansen 1996, 1997; Vamvakas et al. 1996). TEQs estimated on the basis of instrumental analysis do not account for those interactions. In contrast, bioassay-derived TEQs integrate potential interactions among AhR agonists and other compounds by measuring a receptor-mediated response (Tillitt et al. 1991; Sanderson et al. 1998). Comparison of bioassay-derived TEQs with those of TEQs estimated for the same sample by instrumental analysis also suggest that additive and nonadditive interactions exist in biota (Tillitt et al. 1992; Williams and Giesy 1992; Mac et al. 1993). Often those interactions are ignored in the application of TEFs to risk assessment. Ignoring nonadditive interactions in the TEF approach has been justified because (1) the antagonistic or synergistic effects are observed at only very high doses, and the magnitude of those interactions are smaller than the uncertainties already present in the TEF values; (2) the observed nonadditive effects are highly species-, response- and dose-dependent, and their relevance might be of minimal importance; and (3) the mechanisms responsible for these nonadditive effects are unknown (Ahlborg et al. 1992). In general, complex mixtures of PCBs are slightly less than additive; therefore, calculating a TEQ based on an additive model is considered to be a conservative approach (i.e., protective).
Specificity of TEFs
The TEF approach assumes that the rank order of relative potencies of congeners are the same among species. There are, however, considerable variations in the potency of mono-ortho and non-ortho PCBs among mammalian, teleost, and avian models. Therefore, the application of TEFs from rodent bioassays for the assessment of risks in aquatic mammals (e.g., dolphins and whales) might not be appropriate. In addition, PCB congeners have different potencies for various endpoints, resulting in a range of potency values from which the congener-specific TEF is derived. Therefore, the predictive ability of the TEF approach is species- and endpoint-dependent (Safe 1994; Metcalf and Haffner 1995; Seed et al. 1995), which can add uncertainties of a few orders of magnitude to the risk assessment. There is also evidence of age- and sex-specific differences in sensitivity to PCB toxicity (Bosveld et al. 1997) that could add uncertainty in the derivation of TEFs.
Many TEFs are based, in part, on a congener’s ability to induce P-450 enzymes. The induction of P-450 enzymes could be an adaptive mechanism and might not necessarily indicate a toxic effect. That induction is also sometimes nonspecific. Therefore, the use of EROD cytochrome P-450 induction (e.g., measuring EROD activity) to measure TEF and, eventually in risk assessment, requires careful interpretation.
Presence of Non-AhR-Mediated Effects
The TEF approach has been validated for estimating the risks of non-ortho-substituted planar PCB congeners that exhibit dioxin-like activities based on their ability to interact with and activate the Ah receptor. The TEF approach does not consider potential adverse effects of ortho-substituted nonplanar PCB congeners that do not interact with the AhR, but elicit nondioxin-like effects (Seegal 1996). Therefore, the use of TEQs for assessing the potential toxicity of PCBs might not address all the issues of potential adverse effects by PCBs, and the risk to humans or wildlife following exposure to complex mixtures of PCBs might be biased or underestimated (Safe 1994; Birnbaum and DeVito 1995; Seegal 1996). That possibility is especially important because only a small portion of the total mass of PCB mixtures are coplanar non-ortho congeners that elicit dioxin-like activities (Birnbaum and DeVito 1995; Safe 1990; Neubert et al. 1992; Neumann 1996). If the dioxin-like PCBs are the critical contaminant, then variation among mixtures can be reduced by the TEQ approach. For example, application of the TEF concept to examine the tumor promotion potential of Aroclor 1260 underpredicted the observed carcinogenic potential, because ortho-substituted PCB congeners, such as 2,2′,4,4′,5,5′-hexachlorobiphenyl (PCB 153), which are major constituents of Aroclor 1260, are potent tumor promoters (Smith 1997). However, if the critical mechanism of action is caused by non-TCDD-like compounds, the use of the TEQ approach would not be accurate.
The effects of non-TCDD-like PCBs are probably not the critical toxic effects of PCBs. First, relatively high concentrations of lesser chlorinated diortho-substituted congeners need to accumulate in the brain to cause the observed effects. Second, field studies suggest that the active congeners do not tend to accumulate in the brains of animals exposed to complex mixtures of PCBs in the environment. Finally, some neurotoxic effects are mediated by the less-chlorinated PCBs, which are more easily degraded in the environment, bioconcentrate less, and are more readily metabolized and excreted.
Toxicokinetics
The TEF values for dioxin-like PCBs have been derived mainly from short-term tests and in vitro assays (Safe 1994). Such studies might not reflect the pharmacokinetics, metabolism, and excretion that affect the concentration of a chemical at the target organ (De Vito et al. 1995; Lawrence and Gobas 1997). When extrapolating across species, the toxicokinetics must be identical, or differences have to be taken into account. Some factors that affect the interspecies differences have been reviewed (Barrett 1995). In addition, species- and tissue-specific differences in the binding properties, specificity, and physical-chemical properties of the AhR, and the contribution of other P-450 genes to HAH-induced activities challenge the generalities of assumptions of the TEF approach.
Exposure of experimental animals to weathered PCBs might provide more realistic estimates for the risk assessment of ortho-substituted PCB congeners. A few studies have examined behavioral alterations in rats following exposure to contaminated fish from the Great Lakes (Hertzler 1990; Daly 1993). Rats fed different amounts of Great Lakes fish (8%, 15%, and 30% of the diet) for 20 days exhibited behavioral alterations, including reduced exploratory activity, and decreased rearing and nose-poke behavior, which were not exhibited by controls. PCB concentrations in the fish were in the range of 4 to 19 μg/g (wet wt), and total PCB concentrations in the brains of the rat after the exposure period was 50–78 ng/g (wet wt). In contrast, in another study, no significant behavioral effects were seen in rats following a 90-day exposure to PCB-contaminated Great Lakes fish, even though accumulation of ortho-substituted congeners, such as 2,2′,4,4′-(PCB 47), 2,2′,5,5′-(PCB 48), 2,2′,4,4′,5,5′-(PCB 153), 2,2′,5,5′-(PCB 52), 2,4,4′,5-(PCB 74), and 2,2′,4,5′-(PCB 49), in the brain was in the range of 2.5 to 18 ng/g (wet wt) (Beattie et al. 1996). However, the presence of several other contaminants in the diet, such as methylmercury, could confound those studies.
Dose-Response Relationships
When developing relative potency values, a number of assumptions are made about the dose responses for the various compounds. One assumption is that the maximum achievable response for the endpoint of interest is identical for the chemicals evaluated and TCDD (i.e., the congener of interest, although less potent, must have the same efficacy as TCDD). It is also assumed that the dose-response curves are parallel and that they have the same origin. Based on both theoretical analyses and empirical examples from some
studies that developed TEFs, it has been demonstrated that these assumptions for dose-response relationships are seldom met (Putzrath 1997). For example, the slopes of the dose-response curves for many endpoints are different (De Vito et al. 1994). It has been suggested that the relative potency among chemicals would be more accurately represented by a function rather than a point estimate, such as the EC50 or LD50 (lethal dose to 50% of the test population), which are generally used to estimate relative potency (Neubert et al. 1992; Putzrath 1997). That could be accomplished by the use of probability functions.
Extrapolation of Dose Ranges and Routes of Exposure
Most of the information used for establishing TEFs has come from in vitro studies of the induction of monooxygenases and, more recently, from subchronic toxicity studies. Many of the in vivo studies examined acute high-dose effects, such as lethality. Chronic, low-level effects are more relevant to real-world scenarios; therefore, TEFs derived from high-dose exposures might be questionable. The use of different dose regimens in toxicity studies adds further uncertainty to the derived TEF value and eventually to the risk-assessment process.
TOXICITY REFERENCE VALUES
A toxicity reference value (TRY) is a concentration of a chemical in water, food, or a tissue4 that is not expected to cause toxicological effects in the organism of concern. Ideally, TRVs are derived from chronic toxicity studies in which an endpoint relevant to ecological systems or humans was assessed in the species of concern or a closely related species. TRVs are usually derived from the no-observed-adverse-effect level (NOAEL) or the lowest-observed-adverse-effect level (LOAEL). Alternatively, TRVs can be expressed as the geometric mean of the NOAEL and LOAEL to provide a conservative estimate of a threshold of effect (Tillitt et al. 1996).
TRVs that are based on data in the species of interest are not available for the majority of wildlife. Therefore, it is often necessary to use experimental data from a surrogate species to derive a TRV. Uncertainties are associated with the extrapolation of laboratory toxicity data to species exposed in the environment, and the magnitude of those uncertainties should be accounted
for in the TRV. First, there is a wide range of sensitivities to AhR-active chemicals among even closely related species (Gasiewicz 1997). Second, most laboratory studies of toxicity are based on exposure to a parent Aroclor mixture. That mixture might be substantially different from the mixture to which animals in the environment are exposed. Third, if a TEF-TEQ approach was used in the derivation of the TRV, uncertainty could result from the appropriateness of the TEFs. For example, Elliott et al. (1996) derived TEQ-based TRVs for bald eagles using a mammal-based set of TEFs, because bird-specific TEFs were not available. If new data are not available with which to calculate the appropriate TEFs, the resulting uncertainty should be accounted for.
Because of those uncertainties, it is essential to evaluate the applicability of the toxicological data on a site-specific basis for the different exposure pathways and organisms of concern. Appropriate UFs should be applied for each scenario. Uncertainty concerning interpretation of the toxicity test information among different species, different laboratory endpoints, and differences in experimental design (e.g., age of test animals and duration of test) are addressed by applying UFs to the toxicology data to derive the final TRV (TetraTech 1998).
Methods for applying UFs have been published (Opresko et al. 1994; EPA 1995). For example, the method published by EPA Region 8 for the Rocky Mountain Arsenal (RMA) (EPA 1997) uses three UFs: intertaxon variability extrapolation, where values range from 1 to 5; exposure duration extrapolation, where values range from 0.75 to 15; and toxicological endpoint extrapolation, where values range from 1 to 15. Modifying factors, which incorporate other sources of uncertainty, are also used, including threatened, or listed, and endangered species, where values range from 0 to 2; relevance of endpoint to ecological health, where values range from 0 to 2; extrapolation from laboratory to field, where values range from −1 to 2; study conducted with relevant co-contaminants, where values range from −1 to 2; endpoint is mechanistically unclear (versus clear), where values range from 0 to 2; study species is either highly sensitive or highly resistant, where values range from −1 to 2; ratios used to estimate whole-body burden from tissue or egg, where values range from 0 to 2; intraspecific variability, where values range from 0 to 2; and other applicable modifiers, where values range from −1 to 2. The TRV is calculated by dividing the NOAEL or LOAEL from the critical experimental study by the product of the UFs and by the sum of the modifying factors.
In addition to dietary and media-specific TRVs for PCBs, tissue-residue-based TRVs are being used increasingly to evaluate the potential for adverse effects due to PCBs. Tissue-residue-based TRVs have been compiled for fish, birds, and mammals. To derive tissue-residue-based TRVs, site-specific
parameters for exposure to upper trophic level species, including concentrations of contaminants in prey, are used to estimate relevant tissue concentrations in the species of concern. It is important to note that the accuracy of this approach depends on the availability of sufficient data to properly verify food-chain exposure models. Tissue-residue-based effect level data are gaining increasing regulatory acceptance as evidenced in the “Canadian Tissue Residue Guidelines (TRG) for Polychlorinated Biphenyls for the Protection of Wildlife Consumers of Aquatic Biota” (CCME 1998).
Fish Effect-Based Data
Adult fish are exposed to PCBs and related compounds via water, sediment, and food. Eggs and embryos can accumulate these highly lipophilic chemicals from the female during vitellogenesis. Bioaccumulation of PCBs by fish is dependent on the physical and chemical characteristics of individual congeners and on the biotransformation and elimination rates of congeners by fish. The log octanol/water partition coefficient (Kow) of PBCs increases with molecular weight from approximately 4.5 to 8.2 (Mackay et al. 1992), indicating that all but the most highly chlorinated congeners are efficiently bioaccumulated. As a result of these factors, fish preferentially bioaccumulate highly chlorinated penta-, hexa-, and hepta-chlorinated biphenyls, but not deca-chlorinated biphenyls (Walker and Peterson 1994).
Early life stages generally represent the most sensitive developmental stage of fish to chemical contaminants. Thus, accumulation of these persistent chemicals during early life stages in fish is critical to the characterization of risks posed by exposure to PCBs. Signs of toxicity and histopathological lesions produced by PCBs and related compounds in juvenile fish are similar to those seen in higher vertebrates and include decreased food intake, wasting syndrome, delayed mortality, and lesions in epithelial and lymphomyeloid tissues. Toxicity and histopathological lesions produced by PCBs during early development in fish are characterized primarily by cardiovascular and circulatory changes, edema, hemorrhages, and mortality (Walker and Peterson 1991). As with higher vertebrates, there is a great variability in species sensitivities of fish to PCBs. Freshwater salmonid species, particularly lake trout and rainbow trout, are the most sensitive of fish species.
The adverse effects of PCBs on fish have been studied by two primary experimental methods: (1) laboratory exposure of fish to single congeners or technical mixtures via water, diet, or intraperitoneal or in ovo injection; or (2) correlation of concentrations of PCBs and related compounds in the environment with abnormalities in fish populations (e.g., mortality during early devel-
opment or thyroid hyperplasia in adult fish (Walker and Peterson 1991). Although field research can integrate the impact of multiple environmental contaminants on fish, laboratory exposures can identify the specific responses associated with exposure to a single toxicant and determine the dose-response relationships for those responses. Thus, laboratory research can elucidate whether the body burden of a particular contaminant in fish in the environment is capable of producing the abnormalities observed in feral fish populations. Both field research and laboratory research are vital to understanding the toxicity of PCBs to fish and in predicting the risk that these compounds pose to fish in the environment. The available in vivo, in ovo, and in vitro toxic responses of fish characterized in laboratory and field studies were evaluated. However, for the purposes of summarizing effect levels in fish, only data from selected studies using the most relevant aquatic organisms were evaluated.
The majority of the information available on the toxicity of PCBs to fish is from laboratory water-exposure experiments; however, the major route of exposure is likely to be via the food chain. Thus, it is difficult to make accurate estimates of risk when the exposure pathways for actual exposure and laboratory studies are different. As an alternative approach, available studies reporting correlations between concentrations of PCBs in tissue and observed effects were evaluated. Recently, Jarvinen and Ankley (1999) compiled such data for a variety of chemicals, including PCBs. With the use of that data set, effect levels were determined for marine fish species in which critical life stages were evaluated for effects, and tissue concentrations were measured (Table G-4). Studies were evaluated for comparability to the potential species of interest, strength of the cause-effect linkage, exposures to critical life stages (embryos, fry, and juveniles), and sensitive developmental toxicity endpoints, including decreased survival and decreased growth. These tissue-residue effect levels incorporate all of the possible fish exposure pathways, including water, diet, and sediment ingestion. Because the data are so limited in number, a NOEL (no-observed-adverse-effect level) and LOEL (lowest-observed-adverse-effect level) could not be readily determined. Thus, a geometric mean of the available data was calculated to estimate a toxicity threshold in tissue (whole-body or fillet). Niimi et al. (1996) presented a similar table with greater tissue-residue-based effect threshold values (Table G-5). To illustrate the relevance of a tissue-based effect threshold and independence of exposure route, Walker and Peterson (1994) measured concentrations of TCDD in the eggs of rainbow trout after exposure by maternal transfer, water uptake, and injection. The results from this study show that the tissue-based effect level is consistent regardless of exposure route (Table G-6).
TABLE G-4 Tissue-based Effect Concentrations of Total PCBs in Early Developmental Stages of Marine Fish
Effect Level/Species |
Life Stage |
Exposure Route |
Duration (Days) |
Whole bodya |
Filleta |
Effect |
Reference |
NOAEL |
|||||||
Sheepshead minnow, Cyprinodon variegatus |
Embryo |
Adult fish; 49 µg/g |
5 |
27 |
9 |
Survival—no effect |
Hansen et al. 1973 |
Sheepshead minnow, Cyprinodon variegatus |
Embryo-larvae |
Adult fish; 1.9–2.5 µg/g |
28 |
0.88 |
0.3 |
Survival—no effect |
Hansen et al. 1973 |
Pinfish, Lagodon rhomboides |
Juvenile |
water; 100/L |
2 |
17 |
6 |
Survival—no effect |
Duke et al. 1970 |
Spot, Leiostomus xanthrus |
Juvenile |
water; μg/L |
33–56 |
27 |
9 |
Survival—no effect |
Hansen et al. 1971 |
LOAEL |
|||||||
Sheepshead minnow, Cyprinodon variegatus |
Embryo |
Adult fish |
170 |
59 |
Reduced survival |
Hansen et |
al. 1973 |
Sheepshead minnow, Cyprinodon variegatus |
Embryo-larvae |
Adult fish; 9.3–9.7 µg/g |
28 |
5.1 |
2 |
Reduced survival |
Hansen et al. 1973 |
Pinfish, Lagodon rhomboides |
Juvenile |
Water; 5 μg/L |
14 |
14 |
5 |
Reduced survival |
Hansen et al. 1971 |
Spot, Leiostomus xanthrus |
Juvenile |
Water; 5 μg/L |
20–26 |
46 |
16 |
Reduced survival |
Hansen et al. 1971 |
GEOMEAN OF ALL VALUES for PCBs |
|
|
9.2 |
3.2 |
|
|
|
aA relationship between whole body and skin-off fillet was utilized to derive fillet concentrations. The units are mg/kg, wet weight. |
TABLE G-5 Summary of PCB Concentrations in Algae, Zooplankton, and Macroinvertebrates at Which Adverse and Chronic Effects, Cytological Changes, and Changes in Biochemical Activity Levels Occur Based on Short-and Long-term Laboratory Studiesa
Concentrations of dissolved PCBs that are toxic to fish also are very low, particularly following chronic exposure. The 96-hr LC50 values for fathead minnows are 8–15 μg/L for different PCB mixtures (Nebecker et al. 1974). The threshold levels for effects of TCDD in 960-hr exposures to fathead minnows vary from 0.0001 μg/L (for retarded growth and development) to 0.01 μg/L for 100% mortality (Helder 1980, 1981). A no-observed-effective concentration for TCDD has been estimated to be less than 38 pg/L in fingerling rainbow trout (one of the most sensitive fish species to TCDD) exposed for 28 days (Mehrle et al. 1988). However, such relatively short-term exposures do not generally address the chronic effects of dioxin-like chemicals, particularly because delayed mortality, usually after several weeks, occurs at lesser concentrations than acute effects.
Aroclor-based Data—Several water exposure and a few dietary toxicity studies have been conducted with Aroclors. Some of the limitations of the Aroclor-based data are that (1) very few include marine species, and (2) as discussed previously, it might be inappropriate to compare laboratory exposures to Aroclors with field exposures to weathered PCB mixtures. Thus, consideration of these limitations should be made to understand the uncertainty associated with estimation of risk based on TRVs from laboratory exposure to technical PCB mixtures.
Total PCB Data—Considerable field data and limited laboratory data are available in which the concentrations of total PCBs in fish have been
TABLE G-6 Effect of Exposure Route on the Lethal Potency of TCDD to Rainbow Trout Eggs
Exposure Route |
NOAEL μg/kg egg |
LOAEL μg/kg egg |
LD50 μg/kg egg |
LD100 μg/kg egg |
Maternal |
0.023 |
0.05 |
0.058 |
0.145 |
Water uptake |
0.034 |
0.04 |
0.069 |
0.119 |
Egg injection |
0.044 |
0.055 |
0.080 |
0.154 |
Source: Data from Walker and Peterson (1994). |
measured and related to particular adverse effects. The main advantage of these studies is that PCBs were quantified in the tissue of several feral species. One of the main limitations of these data is that there are potential cocontaminants that might confound the interpretation of effect levels from field studies for fish.
PCB Congener and TCDD-Equivalent Data—Considerable field and laboratory data (water exposure, dietary, intraperitoneal injection, and egg injection) are available in which concentrations of PCB congeners, TCDD, or TEQs in eggs or other fish tissue have been measured and related to a particular adverse effect. The main advantage of these studies is that there are several laboratory studies that have been conducted under controlled conditions. Some of the limitations of this data are that (1) very few of these studies have been conducted on marine species, and (2) for the field studies, there are potentially co-contaminants that might confound the interpretation of effect levels. (In many cases, however, co-contaminant data are available from these same studies.)
Avian Effect-Based Data
There is a great deal of information available on the toxicity of PCBs to birds compared with other biota. The information includes both dietary and tissue-residue-based effect levels of PCBs. Some toxic effects of PCBs and related compounds are listed in Table G-7. Although several earlier studies with juvenile and adult birds have shown lethal and biochemical effects of PCBs and related compounds, few studies were designed in such a way that TRVs could be determined reliably. Results of earlier studies, conducted before 1996, have been critically examined in Hoffman et al. (1996). Toxic effects of PCBs and related compounds in birds have been studied by in vivo exposure, in ovo exposure by
TABLE G-7 Toxic Effects of PCBs and TCDD-Equivalents Observed in Birds
Embryo lethality Decreased productivity Liver mfo induction Unabsorbed yolk sacs Vitamin A depletion Porphyria Teratogenesis:
|
egg injection, and in vitro exposure with cultured avian hepatocytes. Due to better control of exposure dose and timing, egg injection studies are often of more use in deriving TRVs than in vivo or in vitro exposure studies. In most cases, embryotoxic and teratogenic effects of PCBs seem to be the most sensitive and ecologically relevant endpoints in birds (Hoffman et al. 1998). Thus, results from in ovo studies are particularly relevant for developing tissue-residue-based toxicity thresholds.
Recently, in ovo studies of dioxin-like compounds have been described (Powell et al. 1996a; Hoffman et al. 1998). Several of these studies were well-conducted and feature an adequate number of replications; several sensitive parameters were monitored. Values for LOAEL, NOAEL, and EC50 that are useful for derivation of TRVs have been estimated for certain PCB congeners from in ovo studies. Additionally, for many bird species, the most sensitive dose metric or effects predictor for PCBs and other dioxin-like chemicals is PCB concentrations in eggs rather than adult tissue (Giesy et al. 1994a,b). That result is in part due to the sensitivity of developing embryos of many species (birds, fish, and mammals) and to the relative tolerance of adults to the effects of dioxin-like chemicals. For example, the LD50 for chicken eggs (Henshel et al. 1993) is 200-fold less than the LD50 for an adult chicken on a wet-weight basis (Greig et al. 1973). Furthermore, LOAEL values for developmental toxicity occur at doses that are
approximately 10-fold lower than LD50 endpoints. Thus, when trying to characterize risk to avian species for dioxin-like compounds, the most sensitive endpoint is developmental toxicity.
Available toxicological studies that correlated effects with PCB concentrations in eggs were evaluated (Tables G-8 through G-12). Similar tissue-residue effects thresholds have been summarized elsewhere (Hoffman et al. 1996) (Table G-13). Some of the dietary avian toxicological studies from the 1970s are still the most useful for deriving PCB reference doses. Alternatively, a tissue-residue-based approach can be used in which observed effects are compared with a known dose to the egg or tissue residues of total PCBs (from a congener-specific analysis) or TEQs. The results of egg-injection studies for predicting potential embryotoxicity of PCBs and TCDD compare favorably with those of feeding studies. In studies in which the same chemicals have been administered by both methods, the egg concentrations required to elicit effects are quite similar for both methods. The biological effects of Aroclor mixtures, individual PCB congeners, TCDD, and TCDD equivalents have been assessed with egg-injection experiments. For risk-assessment purposes, it is possible to model concentrations of PCBs in bird eggs by using published biomagnification factors (Table G-14). The collection of site-specific foraging information and determination of PCB congener concentrations in prey and receptor tissues can improve the accuracy of the predicted egg concentrations. The predicted or measured concentrations can then be compared with TRVs to estimate the magnitude of possible risks.
Birds demonstrate considerable differences in species sensitivities to PCBs and related dioxin-like chemicals among species. In particular, chickens, which are the most frequently used species for PCB exposures are among the most sensitive of the avian species to the effects of PCBs and dioxin-like chemicals. Thus, in most cases, it would be inappropriate to use the chicken as a surrogate species when the application of UFs would result in unrealistically low TRVs for the risk assessment of less-sensitive species, such as bald eagles, gulls, and terns. It is recommended that, wherever possible, family-specific, if not species-specific, toxicity data (whether based on dietary or tissue-residue TRVs) be chosen to most closely match the receptor of concern.
A number of studies with hepatocytes prepared from domestic and wild birds have examined the toxicity of PCBs and related compounds. EROD (a mixed function oxygenase enzyme) activity has been the most commonly measured endpoint in these studies. These studies suggest that, in general, EROD induction is not a toxic response per se but an adaptive biochemical response that is associated with exposure and some of the toxic effects of these chemicals. This reaction is catalyzed predominately by CYP1A1 with some contribution from CYP1A2 and CYP1B1. Excessive induction of mixed function oxygenase activity contributes to TCDD toxicity. Studies with cultured chicken embryo hepatocytes indicate that
TABLE G-8 Avian PCB Toxicity Summary for Dietary Exposures to Aroclors
TABLE G-9 Avian PCB Toxicity Summary for Tissue Residue Effect Levels for Aroclors
Species (Study Type) |
Adverse Effects Evaluated |
Congener or Mixture |
Metric (Unit) |
LD64a |
NOAEL |
LOAEL |
References |
Chicken (laboratory) |
Embryo mortality |
A 1242 |
Egg injection (μg/kg egg) |
10,000 |
|
|
Blazak and Marcum 1975 |
Chicken (laboratory) |
Chick growth |
A 1242 |
Tissue (μg/kg egg) |
|
670 |
6,700 |
Gould et al. 1997 |
Mallard (field) |
|
A 1242 |
Tissue (μg/kg egg) |
|
|
105,000 |
Haseltine and Prouty 1980 |
Chicken (laboratory) |
Chick growth |
A 1254 |
Tissue (μg/kg egg) |
|
670 |
6,700 |
Gould et al. 1997 |
Ringed Turtle Dove (laboratory) |
Hatching success |
A 1254 |
Tissue (μg/kg egg) |
|
|
16,000 |
Peakall and Peakall 1973 |
Chicken (laboratory) |
Hatching success |
A 1248 |
Dose (μg/kg day) |
|
490 |
4,900 |
Scott 1977 |
Screech Owls (laboratory) |
Hatching Success |
A 1248 |
Dose (μg/kg day) |
|
410 |
|
McLane and Hughes 1980 |
Chicken (laboratory) |
|
A 1232 |
Dose (μg/kg day) |
|
980 |
|
Lillie et al. 1974 |
Chicken (laboratory) |
Hatching success |
A 1232 |
Dose (μg/kg day) |
|
980 |
9,800 |
Lillie et al. 1974 |
Chicken (laboratory) |
Hatching success |
A 1232 |
Dose (μg/kg day) |
|
2,440 |
4,880 |
Lillie et al. 1975 |
Chicken (laboratory) |
Hatching success |
A 1016 |
Dose (μg/kg day) |
|
2,440 |
|
Lillie et al. 1975 |
aLD64: The dose that is lethal to 64% of a test population. |
non-ortho PCB congeners 126, 81, 77, and 169 are typically the most potent compounds, and the mono-ortho PCBs 66, 70, 105, 118, 122, 156, 157, and 167 and di-ortho PCBs 128, 138, 170, and 180 can also induce EROD activity (Kennedy et al. 1996b). Although in vitro studies can provide sound indications of the relative potency of different congeners, these experiments are of limited value for establishing defensible whole-organism TRVs for individual congeners in birds.
TABLE G-10 Avian PCB Toxicity Summary for Tissue Residue Effect Levels for Total PCBs
Species (Study Type) |
Adverse Effects Evaluated |
NOAELa (Concentration in Egg, μg/kg) |
LOAEL (Concentration in Egg, μg/kg) |
References |
Chicken (laboratory) |
Hatching success |
360 |
2,500 |
Scott 1977 |
Chicken (laboratory) |
Hatching success |
950 |
1,500 |
Britton and Huston 1973 |
Chicken (laboratory) |
Hatching success deformities |
|
4,000 |
Tumasonis et al. 1973 |
Tree Swallow (field) |
Reproductive behavior |
|
5,000–7,000 |
McCarty and Secord 1999 |
Bald Eagle (field) |
Reproductive success |
|
4,000 |
Ludwig et al. 1993 |
Bald Eagle (field) |
Reproductive success |
1300 |
7,200 |
Wiemeyer et al. 1984 |
Bald Eagle (field) |
Reproductive success |
|
13,000 |
Bosveld and van den Berg 1994 |
Bald Eagle (field) |
Hatching success |
400 |
4,000 |
180, source document |
Double-crested Cormorant (field) |
Hatching success |
350 |
3,500 |
23, 24, source document |
Common tern (field) |
Reproductive success |
7000 |
8,000 |
Bosveld and van den Berg 1994 |
Common tern (laboratory) |
Hatching success, deformities |
4800 |
10,000 |
Hoffman et al. 1993 |
Common tern (field) |
Hatching success |
5,200–5,600 |
7,000 |
Becker et al. 1993 |
Forster’s tern (both) |
Hatching success |
4500 |
22,000a |
Kubiak et al. 1989 |
Forster’s tern (field) |
Reproductive success |
7000 |
19,000 |
Bosveld and van den Berg 1994 |
Caspian terns (field) |
Hatching success, deformities |
420 |
4,200 |
Yamashita et al. 1993; Giesy et al. 1994a |
Herring gulls (field) |
Hatching success deformities |
500 |
|
Weseloh et al. 1994; Giesy et al. 1994a |
aNOAEL=2.2 μg/kg TEQ in egg. |
TABLE G-11 Avian PCB Toxicity Summary for Tissue Residue Effect Levels for PCB Congeners
Species |
Adverse Effects Evaluated |
Congener or mixture |
Metric |
LD50a (μg/kg/egg) |
NOAEL (μg/kg/egg) |
LOAEL (μg/kg/egg) |
TEQ LD50 |
References |
Chicken |
Hatching success |
PCB 77 |
Tissue |
8.6 |
|
|
0.43 |
Brunstrom and Andersson 1988 |
Chicken |
|
PCB 77 |
Egg injection |
|
|
30 |
|
Nikolaidis et al. 1988 |
Chicken |
|
PCB 77 |
Egg injection |
2.6 |
0.12 |
1.2 |
|
Hoffman et al. 1998 |
Chicken |
Decreased hatch weight |
PCB 77 |
Egg injection |
8.8 |
1 |
3 |
|
Powell et al. 1996a |
Chicken |
Embryo mortality |
PCB 77 |
Egg injection |
40 |
|
|
|
van den Berg et al 1998 |
Turkey |
Hatching success |
PCB 77 |
Tissue |
~800 |
|
|
|
Brunstrom and Lund 1988 |
Ring-necked pheasant |
Hatching success |
PCB 77 |
Tissue |
|
100 |
|
|
Brunstrom and Reutergardh 1986 |
American kestrel |
Embryo mortality |
PCB 77 |
Egg Injection |
316 |
|
100 |
|
Hoffman et al. 1998 |
Goldeneye |
Hatching success |
PCB 77 |
Tissue |
>1,000 |
|
|
>50 |
Brunstrom and Reutergardh 1986 |
Domestic |
Hatching success |
PCB 77 |
Tissue |
>1,000 |
|
|
>50 |
Brunstrom 1988 |
Mallard |
Hatching success |
PCB 77 |
Tissue |
>5,000 |
|
|
>250 |
Brunstrom 1988 |
Blackheaded gull |
Hatching success |
PCB 77 |
Tissue |
<1,000 |
|
|
<50 |
Brunstrom 1988 |
Herring gull |
Hatching success |
PCB 77 |
Tissue |
>1,000 |
|
|
36892 |
Brunstrom 1988 |
Species |
Adverse Effects Evaluated |
Congener or mixture |
Metric |
LD50a (μg/kg/egg) |
NOAEL (μg/kg/egg) |
LOAEL (μg/kg/egg) |
TEQ LD50 |
References |
Chicken |
Hatching success |
PCB 126 |
Tissue |
|
|
3.2 |
0.32 |
Brunstrom and Andersson 1988 |
Chicken |
Hatching success |
PCB 126 |
Tissue |
2.3 |
|
|
0.23 |
Powell et al. 1996b |
Chicken |
Hatching success |
PCB 126 |
Tissue |
0.4 |
|
|
0.04 |
Hoffman et al. 1998 |
Chicken |
Reproductive behavior |
PCB 126 |
Egg injection |
|
0.5 |
1 |
|
Zhao et al. 1997 |
Chicken |
Deformities; decreased hatch weight |
PCB 126 |
Egg injection |
0.4 |
|
0.3 |
|
Hoffman et al. 1998 |
Bobwhite |
Deformities |
PCB 126 |
Tissue |
24 |
|
|
2.4 |
Hoffman et al. 1998 |
American kestrel |
Deformities |
PCB 126 |
Egg injection |
65 |
2.3 |
23 |
6.5 |
Hoffman et al. 1998 |
Dbl-crested cormorant |
Hatching success |
PCB 126 |
Tissue |
158 |
|
|
16 |
Powell et al. 1997b |
Dbl-crested cormorant |
Embryo mortality |
PCB 126 |
Egg injection |
|
200 |
400 |
|
Powell et al. 1997a |
Common tern |
Hatching success; Embryo mortality |
PCB 126 |
Egg injection |
104 |
|
44 |
10.4 |
Hoffman et al. 1998 |
Common tern |
Embryo mortality |
PCB 126 |
Egg injection |
45 |
|
|
|
Hoffman et al. 1998 |
Chicken |
Embryo mortality |
PCB 157 |
Egg injection |
2,000 |
|
|
|
van den Berg et al 1998 |
TABLE G-12 Avian PCB Toxicity Summary for Dietary Exposures and Tissue Residue Effect Levels for TCDD and TCDD Equivalents
Species (Study Type) |
Adverse Effects Evaluated |
Congener or Mixture |
Metric (μg/kg/egg) |
LD50a |
NOAEL |
LOAEL |
References |
Ring-neck pheasants (laboratory) |
Embryo mortality |
TCDD |
Dose |
|
0.014 |
0.14 |
Nosek et al. 1992a,b, 1993 |
Chicken (laboratory) |
Embryo mortality |
TCDD |
Egg injection |
0.115 |
|
|
Henshel et al. 1993 |
Chicken (laboratory) |
Embryo mortality |
TCDD |
Egg injection |
0.18 |
|
|
Henshel et al. 1993 |
Chicken (laboratory) |
Embryo mortality |
TCDD |
Egg injection |
0.24 |
|
|
Allred and Strange 1977 |
Chicken (laboratory) |
Embryo mortality |
TCDD |
Egg injection |
LD100 1.0 |
|
|
Higginbotham et al. 1968 |
Chicken (laboratory) |
Embryo mortality |
TCDD |
Egg injection |
0.15 |
0.08 |
0.16 |
Powell et al. 1996b |
Chicken (laboratory) |
Decreased hatch weight |
TCDD |
Egg injection |
|
0.06 |
0.01 |
Henshel et al. 1997a |
Chicken (laboratory) |
Decreased hatch weight |
TCDD |
Egg injection |
|
0.1 |
0.3 |
Henshel et al. 1997a |
Chicken (laboratory) |
Deformities |
TCDD |
Egg injection |
|
|
0.32 |
Walker et al. 1997 |
Chicken (laboratory) |
Deformities |
TCDD |
Egg injection |
|
0.001 |
0.01 |
Henshel et al. 1997b |
Chicken (laboratory) |
|
TEQ |
Tissue |
LD100 1.0 |
|
|
Giesy et al. 1994a |
Chicken (laboratory) |
|
TEQ |
Tissue |
0.14 |
|
|
Cheung et al. 1981; Giesy et al. 1994a |
Chicken (laboratory) |
Deformities |
TEQ |
Tissue |
0.65 |
|
|
Giesy et al. 1994a |
Chicken (laboratory) |
Deformities |
TEQ |
Tissue |
|
|
0.0064 |
Giesy et al. 1994a |
Chicken (laboratory) |
Embryo mortality |
TEQ |
Tissue |
LD100 1.0 |
|
|
Higgenbotham et al. 1968; Giesy et al. 1994a |
Pheasant (laboratory) |
Embryo mortality |
TEQ |
Egg injection |
1.4–2.2 |
|
|
Nosek et al. 1993 |
Wood duck (field) |
Reproductive success |
TEQ |
Tissue |
|
|
0.02 |
White and Hoffman 1995; Giesy et al. 1994a |
Wood duck (field) |
Hatching success |
TEQ |
Tissue |
|
£5 |
>20–50 |
White and Hoffman 1995 |
Species (Study Type) |
Adverse Effects Evaluated |
Congener or Mixture |
Metric (μg/kg/egg) |
LD50a |
NOAEL |
LOAEL |
References |
Dbl-crested cormorant (laboratory) |
Embryo mortality |
TCDD |
Egg injection |
|
1 |
4 |
Powell et al. 1997a |
Dbl-crested cormorant (field) |
Hatching success |
TEQ |
Tissue |
~0.55 |
|
|
Tillet et al. 1992 |
Dbl-crested cormorant (field) |
Hatching success |
TEQ |
Tissue |
LD100 1.03 |
|
|
Giesy et al. 1994a |
Dbl-crested cormorant (field) |
Hatching success |
TEQ |
Tissue |
LD37 0.344 |
|
|
Giesy et al. 1994a |
Dbl-crested cormorant (field) |
Hatching success |
TEQ |
Tissue |
LD27 0.217 |
|
|
Giesy et al. 1994a |
Dbl-crested cormorant (field) |
Hatching success |
TEQ |
Tissue |
LD8 0.35 |
|
|
Giesy et al. 1994a |
Dbl-crested cormorant (field) |
Embryo mortality |
TEQ |
Tissue |
0.46 |
|
|
Tillitt 1989 |
Dbl-crested cormorant (field) |
Hatching success |
TEQ |
Tissue |
0.46 |
|
|
Tillitt et al. 1992; Giesy et al. 1994a |
Caspian tern (field) |
Hatching success |
TEQ |
Tissue |
0.75 |
|
|
Giesy et al. 1994a |
Common tern (laboratory) |
Hatching success |
TEQ |
Tissue |
|
<1 |
|
Bosveld and van den Berg 1994 |
Herring gull (field) |
Hatching success |
TEQ |
Tissue |
|
36892 |
|
Ludwig et al. 1993 |
Herring gull (field) |
Hatching success |
TEQ |
Tissue |
LD19 0.557 |
|
|
Giesy et al. 1994a |
Osprey (field) |
Reproductive success |
TEQ |
Tissue |
0.14 |
|
|
Woodford et al. 1998 |
Bald eagle (field) |
Reproductive success |
TEQ |
|
0.2 |
|
|
Elliott et al. 1996 |
Great blue heron (field) |
Deformities; Chick growth |
TEQ |
|
|
0.02 |
0.245 |
Hart et al. 1991 |
aLD50: The dose that is lethal to 50% of a test population. |
TABLE G-13 Summary of PCB and TCDD Threshold Effect Levels in Birds
Concentration |
Effect |
20 to 50 ppt of TCDD in eggs |
Embryo mortality and teratogenesis in chickens, decreased productivity and teratogenesis for wood ducks |
150 to 250 ppt of TCDD in eggs |
Decreased embryonic growth, edema in herons |
618 to 7,366 ppt of TCDD equivalents (congener chemistry) |
Embryotoxicity in Forester’s tern |
1,000 ppt of TCDD in eggs |
Embryo mortality in pheasants |
1 to 5 ppm of total PCBs in eggs |
Decreased hatching success for chickens |
8 to 25 ppm of total PCBs in eggs |
Decreased hatching success for terns, cormorants, doves, eagles |
Source: Data from Hoffman et al. (1996). |
TABLE G-14 Biomagnification Factors from Alewife to Herring Gull Egg for Dioxin-like Compounds
Compound |
Biomagnification Factor |
2,3,7,8-TCDD |
21 |
Total PCBs |
32 |
2,3,3′,4,4′-Penta-CB (105) |
20 |
2,3,4,4′,5-Penta-CB (114) |
31 |
2,2′,3,4,4′,5-Hexa-CB (138) |
42 |
3,3′,4-Tri-CB (35) |
0.8 |
3,3′,4,4′-Tetra-CB (77) |
1.8 |
3,3′,4,4′,5-Penta-CB (126) |
29 |
3,3′,4,4′,5,5′-Hexa-CB (169) |
46 |
Source: Data from Hoffman et al. (1996). |
In addition, in vitro studies do not account for pharmacokinetic and pharmacodynamic parameters, which could alter the toxic potential of a congener.
For a tissue-residue-based TRV, several studies for wildlife have been evaluated. A specific recommendation cannot be made at this point, because the TRVs are usually species-specific and thus should be selected based on similarity to
receptors of concern. It is, therefore, essential to perform a critical evaluation of the applicability of the toxicological data to the site-specific receptors of concern and exposure pathways. TRVs derived in the same species are not available for the majority of wildlife receptors and, therefore, it is necessary to derive TRVs using toxicological data for surrogate species in combination with UFs.
Tissue-residue-based TRVs are being used increasingly to evaluate the potential for adverse effects due to PCBs. For the purposes of this report, the term “tissue-residue-based TRV” is synonymous with “maximum allowable tissue concentration (MATC),” a term that is sometimes used by agencies and reported in the literature. To derive tissue-residue-based TRVs, site-specific parameters for exposure to upper trophic level receptors, including concentrations of contaminants in prey, are used to estimate relevant tissue concentrations in the receptors of concern. It is important to note that the accuracy of this approach depends on the availability of sufficient data to properly verify food-chain exposure models. Tissue-residue-based effect level data are gaining increasing regulatory acceptance as evidenced in the “Canadian Tissue Residue Guidelines (TRG) for Polychlorinated Biphenyls for the Protection of Wildlife Consumers of Aquatic Biota” (CCME 1998).
When deriving TRVs for a particular receptor species, there are a number of uncertainties that need to be taken into account. One way to do that is to assign UFs to assure that the derived TRV is protective of the receptor of interest. Thus, it is important to remember that the derivation and application of TRVs is part of a conservative process, meant to be protective rather than predictive. For this reason, a tiered risk-assessment approach that allows for more and more refined estimates of TRVs needs to be derived. Each level of assessment requires more sophisticated estimates of both exposure and response. The TRV provides information only on the response parameters. Refining TRV estimates might require additional toxicity testing either for a particular mechanism of action, species, or vector of exposure.
For bird species, the most sensitive dose metric for dioxin-like chemicals are egg concentrations rather than adult tissue concentrations (Giesy et al. 1994a,b). That is in part due to stage-specific sensitivity of many species (birds, fish, and mammals) during development and relative tolerance of adults to the effects of dioxin-like chemicals. In other words, effects can be seen at lesser doses for young, developing animals than for adults. For example, the LD50 for chicken eggs (Henshel et al. 1993) is 200-fold less than the LD50 for an adult chicken (on a wet-weight basis) (Greig et al. 1973). Furthermore, LOAEL values for developmental toxicity occur at doses that are approximately 10-fold lower than LD50 endpoints. Thus, when trying to characterize risk to avian species for dioxin-like compounds, the most sensitive endpoint is developmental toxicity.
The biological effects of Aroclor mixtures, individual PCB congeners, TCDF, and TCDD equivalents have been assessed with egg-injection experiments. For
risk-assessment purposes, it is possible to model concentrations of PCBs in bird eggs using published biomagnification factors. The collection of site-specific foraging information and determination of PCB congener concentrations in prey and receptor tissues can improve the accuracy of the predicted egg concentrations. The predicted or measured concentration can then be compared with TRVs to estimate the magnitude of possible risks.
Aroclor-Based Data—Some of the most frequently cited bird studies for controlled, laboratory, dietary exposures are Aroclor-based exposure studies (Table G-8). In particular, the pheasant study of Dahlgren et al. (1972) and the chicken study of Platonow and Reinhart (1973) are often selected for development of TRVs. The Great Lakes Water Quality Initiative report (EPA 1995a) used the pheasant study of Dahlgren et al., (1972) because it is a wildlife species and because the study evaluated a critical life stage. However, to compare estimated exposures for receptors in a risk assessment with this type of benchmark, an estimate of total PCBs from a PCB congener-specific analysis, rather than an Aroclor-based method, is recommended because of environmental weathering processes. Furthermore, although the predominant route of exposure for toxicity studies for Aroclors is dietary, few of the available dietary studies have evaluated the concentration of PCBs in eggs during and after the exposure (Table G-9). Some of the limitations of this Aroclor-based data are that (1) very few of these studies have been conducted on wildlife species, and (2) as discussed previously, it might not be appropriate to compare laboratory exposures to technical Aroclors (with potential contamination by other more potent dioxin-like chemicals) with field exposures to weathered PCBs. These factors need to be considered when estimating TRVs and the uncertainty associated with them.
Total PCB Data—There are many field and laboratory studies in which concentrations of total PCBs in eggs have been measured and related to adverse effects (Table G-10). The main advantages of these studies, especially those conducted in the last 10 years, are that the total weathered PCBs (often by PCB congener analysis as either total PCBs or TEQs) are measured in the tissue of wildlife species and that these concentrations were related to ecologically relevant endpoints. Some of the limitations of this total PCB data are that (1) in a few cases, the individual PCB congeners were not quantified (studies from the 1970s), and (2) there are potential co-contaminants that might confound the interpretation of effect levels from field studies. (Data on some co-contaminants are available from these studies.)
PCB Congener and TCDD-Equivalent Data—Considerable field and laboratory data are available in which the concentrations of PCB congeners, TCDD, or TEQs
in eggs have been measured and related to adverse effects (Table G-11). The main advantages of these studies are that there are several laboratory studies that have been conducted under controlled conditions, and many of the available studies have been conducted with wildlife species. Some of the limitations of these data are that (1) there are no dietary exposures, and (2) for the field studies, there are potential co-contaminants that might confound the interpretation of effect levels. (In many cases, co-contaminant data are available from these studies.)
Aquatic Mammal Effect-Based Data
Aroclor-based Data—No data are available for Aroclor-based toxicological studies.
Total PCB Data—The evaluation of the effects of environmental contamination on the health of aquatic mammals represents a considerable challenge. Logistical considerations make the sampling of large numbers of wild animals difficult, and ethical concerns discourage in vivo studies on captive animals. A few semi-field studies involving the exposure of seals to contaminated fish were used to derive TRVs for PCBs (de Swart et al. 1994; Ross et al. 1995). Immunotoxic endpoints, such as natural killer cell activity in the blood, vitamin A levels, and reproductive effects were measured in these studies. The TRVs derived for exposure studies with seals were compared with those reported for otter (Smit et al. 1996) and mink (Leonards et al. 1995; Tillitt et al. 1996). Because a majority of reports of PCBs in aquatic mammals have been from blubber samples, a blubber-based TRV can be derived by lipid normalizing the blood or liver-based PCB concentrations. Although the results of semi-field studies have been confounded by the occurrence of co-contaminants in the diet and by the lack of dose-response relationships, these studies have suggested that PCBs were the main cause for the observed toxic effects. Furthermore, the exposures mimicked observed field situations and, therefore, these values were considered suitable for deriving a toxicity benchmark to assess risks of PCBs to aquatic mammals. The lipid normalized NOAEL value in seal blood of 11 μg/g was comparable to the otter liver NOAEL of 9 μg/g (Kannan et al. 2000; Table G-15). After reviewing all of the available literature, a tissue-residue NOAEL for PCBs in aquatic mammals of 10 μg/g, lipid wt, has been suggested (Kannan et al. 2000) (Table G-15).
PCB Congener and TCDD-Equivalent Data—Based on immunotoxicological studies in seals, a LOAEL for blubber TEQ of 210 pg/g, lipid wt, has been suggested (Ross et al. 1995). However, the lipid normalized liver TEQs for otters and mink have been estimated to be in the range of 400–2,000 pg/g. That estimate
TABLE G-15 Summary of NOAEL and LOAEL for Total PCBs and TEQs in Semi-Field Investigations with Seals, Dolphins and European Otter
Exposure |
PCBs |
TEQs |
Seals and Dolphins |
||
Daily dose NOAEL |
5.2 μg/kg bw/d |
0.58 ng/kg bw/d |
Daily dose LOAEL |
28.9 μg/kg bw/d |
5.8 ng/kg bw/d |
Dietary NOAEL |
100 ng/g, wet wt |
NA |
Dietary LOAEL |
200 ng/g, wet wt |
NA |
Seal blood NOAEL (lipid 0.05–0.32%) |
5.2 μg/kg, lipid wt |
NA |
Seal blood LOAEL |
25 μg/kg, lipid wt |
NA |
Seal blubber NOAEL |
NA |
90 pg/g, lipid wt |
Seal blubber LOAEL |
NA |
286 pg/g, lipid wt |
Dolphin blood (in vitro) |
26 ng/g, wet wt |
NA |
European Otter |
||
Dietary NOAEL (lipid 6.2%) |
12 ng/g, wet wt (or) 200 ng/g, lipid wt |
1 pg/g, wet wt or 16 pg/g, lipid wt |
Dietary LOAEL |
33 ng/g, wet wt (or) 530 ng/g, lipid wt |
2 pg/g, wet wt 33 pg/g, lipid wt |
Otter liver NOAEL |
170 ng/g, wet wt (or) 4 μg/g, lipid wt |
42 pg/g, wet wt 1 ng/g, lipid wt |
Otter liver LOAEL |
460 ng/g, wet wt (or) 11 μg/g, lipid wt |
84 pg/g, wet wt or 2 ng/g, lipid wt |
implies that sensitivity to TEQs might be greater in seals than in mink or otter. Thus, TRVs for TEQs might be dependent on the aquatic mammal species in question.
Uncertainties
This discussion of uncertainty is designed to assist in the understanding of the relative degree of confidence in the toxicity benchmarks and available data. An uncertainty analysis is required for ecological risk assessments under EPA guidance and should be performed for the quantitative and qualitative parameters that
are included in risk assessments. When appropriate, sensitivity analyses should be performed to illustrate how the results would change if different toxicity benchmarks and other assumptions were incorporated into a particular analyses. Some common factors that contribute to uncertainty include assumptions relating to exposure models and toxicity thresholds. Additional parameters that contribute to uncertainty and their associated concerns include the following:
-
Spatial distribution of contaminants. (What is the exposure from hot spots compared to more diffuse concentrations of contaminants?)
-
Temporal (including seasonal) distribution of contaminants. (Is there temporal variability for exposure pathways? Is there any evidence of natural attenuation and/or a reduction in source inputs?)
-
Bioavailability issues. (Are there factors that cause variability in the bioavailability of PCBs?)
-
Co-contaminants and unknown interactions. (Are there other contaminants besides PCBs or a noncontaminant stressor that is predicted to cause unacceptable risk to the receptors of concern?)
-
Species-specific differences in TEFs.
CONCLUSIONS AND RECOMMENDATIONS
In this appendix, available toxicological data for PCBs derived from dietary or media-specific exposures and tissue residues were evaluated and discussed. A database of studies is available; however, for each taxonomic group, evaluations must be made about the appropriateness and usefulness of data for risk assessments, especially for ecological risk assessments for which the database is more limited. Recommendations for each taxonomic group have been provided separately. In general, a weight-of-evidence approach can be utilized in which multiple measurement endpoint approaches (dietary TRVs, tissue-residue-based TRVs, and field studies) provide separate lines of evidence.
General recommendations are the following:
-
An estimate of total PCBs (from a PCB congener-specific analysis) is sufficient to characterize potential risk to invertebrates. Individual PCB congener data or total PCB data can then be used, if necessary, as input to dietary food-chain models for biota that consume invertebrates.
-
An estimate of tissue residues of total PCBs (from a congener-specific analysis) or TEQs is sufficient to characterize risk to fish populations. Water-based concentration effect levels are not appropriate to evaluate the potential risk
-
of PCBs to fish, because the predominant exposure pathway is not from the water but from dietary sources. Individual PCB congener data or total PCB data can then be used, if necessary, as input to dietary food-chain models for biota that consume fish.
-
An estimate of total PCBs (from a congener-specific analysis) or TEQs is sufficient to characterize risk to birds. The dietary exposure TRV, tissue-residue TRV, or both can be used to characterize risk.
-
An estimate of tissue residues of total PCBs (from a congener-specific analysis) or TEQs is sufficient to characterize risk to marine mammals.
Recommendations for Selection of Toxicity Reference Values
TRVs are ideally derived from chronic toxicity studies in which an ecologically relevant endpoint was assessed in the species of concern, or a closely related species. In this section, available toxicological data for PCBs derived from dietary or media-specific exposures and tissue residues were evaluated and discussed. A database of thousands of studies is available. However, for each group of biota, the appropriateness and usefulness of data must be considered for ecological risk assessments. Thus, for each group of biota, recommendations have been provided separately. In general, a weight-of-evidence approach should be utilized in which multiple measurement endpoint approaches (dietary TRVs, tissue-residue-based TRVs, and field studies) provide separate lines of evidence.
Benthic Invertebrate and Other Lower Trophic Level Biota
Generally, PCBs and other dioxin-like chemicals are not particularly toxic to lower trophic level biota, including algae, zooplankton, and invertebrates as these species lack the cellular Ah receptor that is required to mediate toxicity. It is recommended that an estimate of total PCBs (from a PCB congener-specific analysis) is sufficient to characterize potential risk to invertebrates. Individual PCB congener data or total PCB data can then be used, if necessary, as input to dietary food chain models for biota that consume invertebrates.
Fish
It is recommended that an estimate of tissue residues of total PCBs (from a congener-specific analysis) or TEQs is sufficient to characterize risk to fish. Water-based concentration effect levels are not appropriate to evaluate the poten-
tial risk from PCBs to fish because the predominant exposure pathway is not from the water but from dietary sources. Individual PCB congener data or total PCB data can then be used, if necessary, as input to dietary food chain models for biota that consume fish. To illustrate the relevance of a tissue-based effect threshold and independence of exposure route, Walker and Peterson, (1994) measured concentrations of TCDD in the eggs of rainbow trout after exposure by maternal transfer, water uptake, and injection. The results from this study show that the tissue-based effect level is consistent regardless of exposure route. A geometric mean of the available NOAEC and LOAEC data for marine species of fish was calculated (9.2 mg/kg) to estimate a toxicity threshold in tissue (whole body) based on survival of juveniles (a sensitive life stage).
Birds
For birds, it is recommended that a combination of dietary exposure modeling and tissue-residue-based effect levels be used. The potential for exposure is greatest for top-level predators; thus, several lines of evidence should be evaluated (possibly in a phased approach).
Considerable field and laboratory data are available in which the concentrations of total PCBs, PCB congeners, TCDD, or TEQs in eggs have been measured and related to adverse effects. The main advantages of these studies, especially those conducted in the last 10 years, are that the total weathered PCBs (often by PCB congener analysis as either total PCBs or TEQs) in the tissue of wildlife species was measured and that these concentrations were related to ecologically relevant endpoints.
Very few studies have been conducted from which a dietary TRV for PCB exposure to birds can be derived. Attempts to derive TRVs for birds at Navy sites in the San Francisco Bay have focused on a dietary study by Platonow and Reinhart (1973) (Tetra Tech 1998). However, there appears to be confusion regarding the calculation of the daily intake rate from this study. In different parts of the same report (Tetra Tech 1998) and in an EPA (1995) document, four different NOAEL and LOAEL values have been reported (Table G-16). Because the 1995 EPA document states clearly how its value was calculated, it appears to have the most support. However, for a dietary exposure-based TRV, it is recommended that the study by Dahlgren et al. (1972) be used, because it was conducted on a wildlife species and it evaluated a sensitive life stage. This study was also selected by EPA (1995) to provide a basis for water-quality values protective of wildlife. The NOEL and LOEL values from this study are 0.18 and 1.8 mg/kg/d, respectively.
TABLE G-16 Avian TRVs for PCBs
Original Study |
NOAEL (mg/kg/d) |
LOAEL (mg/kg/d) |
TRV Citation |
Platonow and Reinhart 1973 |
0.09 |
0.88 |
TetraTech 1998, pp. 5–68 |
Platonow and Reinhart 1973 |
0.034 |
0.34 |
TetraTech 1998, p. D-67 |
Platonow and Reinhart 1973 |
0.244 |
2.44 |
EPA (1995) |
Dahlgren et al. 1972 |
0.18 |
1.8 |
EPA (1995 |
For a tissue-residue-based TRV, several studies for wildlife have been evaluated (refer to Chapter 6). A specific recommendation cannot be made at this point, because the TRVs are usually species-specific and thus should be selected on the basis of similarity to the receptors of concern. For example, to be protective of the embryo mortality for double-crested cormorants, a tissue-residue-based TRV would be between 350 (NOAEL) and 3,500 μg/kg (LOAEL) for total PCBs, respectively, and between 1 (NOAEL) and 4 μg/kg (LOAEL) for TEQs, respectively (Giesy et al. 1994a,b).
Aquatic Mammals
It is recommended that a tissue-residue threshold effect concentration for PCBs in marine mammals be 11 μg/g, lipid wt (Kannan et al. 2000) to be protective of immune function.
PCB Congener and TCDD-Equivalent Data—Based on immunotoxicological studies in seals, a threshold effect concentration for blubber TEQ of 520 pg/g, lipid wt, has been suggested (Ross et al. 1995).
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