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-



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A Risk-Management Strategy for PCB-Contaminated Sediments 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-

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A Risk-Management Strategy for PCB-Contaminated Sediments 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.,

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A Risk-Management Strategy for PCB-Contaminated Sediments 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.

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A Risk-Management Strategy for PCB-Contaminated Sediments 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-

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A Risk-Management Strategy for PCB-Contaminated Sediments 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). 1   Some evidence suggests that effects on the dopaminergic system might be involved in the neurotoxicity of PCBs. 2   And 3,3′,4,4′-5-(PCB 12b) (Angus and Contreras 1996).

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A Risk-Management Strategy for PCB-Contaminated Sediments 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

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A Risk-Management Strategy for PCB-Contaminated Sediments 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

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A Risk-Management Strategy for PCB-Contaminated Sediments 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

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A Risk-Management Strategy for PCB-Contaminated Sediments 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

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A Risk-Management Strategy for PCB-Contaminated Sediments 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

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A Risk-Management Strategy for PCB-Contaminated Sediments 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

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