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5—
Effects on Reproduction and Development

The harmful effects of exposure to environmental contaminants on reproduction and development in wildlife populations have been reported in the scientific literature for many years. Reported reproductive disorders in wildlife have included morphologic abnormalities, eggshell thinning, population declines, impaired viability of offspring, altered hormone concentrations. and changes in sociosexual behavior.

Laboratory experiments replicating the adverse effects of exposure to the potent synthetic estrogen diethylstilbestrol (DES) during critical periods in development (Newbold 1995; also see Appendix) have focused attention on the potential of chemicals with estrogenic properties to cause developmental and reproductive hazards.

The adverse consequences of prenatal exposure to DES on the female genital tract in humans have been reviewed in detail by Herbst and Bern ( 1981 ) and by Mittendorf (1995); they are the subject of continued, intensive investigation. Whether exposure to environmental hormonally active agents (HAAs) affects animals and humans similarly is not clear, but because exposure of animals to DES causes alterations in male and female offspring, the possibility must be considered that there will be adverse effects from exposure to other compounds with estrogenic, antiestrogenic, or antiandrogenic activity. There are also concerns that exposure to low doses of certain chemicals at critical stages in organ development can result in abnormalities that lead to irreversible changes in the functioning of organ systems later in life. Such damage would not occur through genetic mutations, but by processes that regulate genes during development and cell differentiation. The effects of hormones in adults are usually transient, and hormonal effects disappear when the chemical is not present. By contrast, environmental chemicals that alter gene activity during development would producecontinue



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Page 119 5— Effects on Reproduction and Development The harmful effects of exposure to environmental contaminants on reproduction and development in wildlife populations have been reported in the scientific literature for many years. Reported reproductive disorders in wildlife have included morphologic abnormalities, eggshell thinning, population declines, impaired viability of offspring, altered hormone concentrations. and changes in sociosexual behavior. Laboratory experiments replicating the adverse effects of exposure to the potent synthetic estrogen diethylstilbestrol (DES) during critical periods in development (Newbold 1995; also see Appendix) have focused attention on the potential of chemicals with estrogenic properties to cause developmental and reproductive hazards. The adverse consequences of prenatal exposure to DES on the female genital tract in humans have been reviewed in detail by Herbst and Bern ( 1981 ) and by Mittendorf (1995); they are the subject of continued, intensive investigation. Whether exposure to environmental hormonally active agents (HAAs) affects animals and humans similarly is not clear, but because exposure of animals to DES causes alterations in male and female offspring, the possibility must be considered that there will be adverse effects from exposure to other compounds with estrogenic, antiestrogenic, or antiandrogenic activity. There are also concerns that exposure to low doses of certain chemicals at critical stages in organ development can result in abnormalities that lead to irreversible changes in the functioning of organ systems later in life. Such damage would not occur through genetic mutations, but by processes that regulate genes during development and cell differentiation. The effects of hormones in adults are usually transient, and hormonal effects disappear when the chemical is not present. By contrast, environmental chemicals that alter gene activity during development would producecontinue

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Page 120 effects much harder or impossible to reverse. Evaluating the effects of such chemicals is more difficult than evaluating the effects of chemicals on adults. Effective doses may be lower than effective doses in adults, and the effects are considerably removed in time from the exposure, which can make causal relationships more difficult to establish. This chapter is a critical analysis of the literature on the link between exposure to HAAs and reproductive and developmental effects observed in laboratory studies, in humans, and in wildlife populations. Only a few (primarily estrogenic) HAAs are covered in this chapter, and evaluations of wildlife are limited to only a few vertebrate species. Although endocrine systems are remarkably well conserved among vertebrate groups, there are significant differences in their operation. While gonadal reversal does not occur in mammals, data from Seveso (Mocarelli et al. 1996) and occupationally exposed cohorts (Goldsmith et al. 1984: Potashnik et al. 1984; Potashnik and Porath 1995) suggest that some selective process may be involved to alter the sex ratio at birth. Hormonal control of sex differentiation is different in birds than it is in mammals, even though the same hormones (for example, estradiol) are involved. There are other important differences in the development of the reproductive systems of various vertebrate groups. Thus, HAAs could affect different vertebrate groups in different ways. For this reason, it is necessary to understand not only the effects of exposure to different HAAs on vertebrate development and reproduction but also the various effects of exposure to single HAAs on reproduction and development in different vertebrate groups. Laboratory studies are discussed for specific HAAs, including some chemicals known to bind to estrogen receptors: insecticides (dichlorodiphenyl-trichloroethane (o,p'-DDT), methoxychlor, and chlordecone); a monomer used in plastic (bisphenol A); an alkylphenol surfactant used in detergents, cosmetics and toiletries, and other household products (octylphenol); and a plasticizer (butyl benzyl phthalate (BBP)). Other compounds are known to bind to androgen receptors: the fungicide vinclozolin and 1,1 -dichloro-2,2-bis(p-chlorophenyl)ethylene (p,p'-DDE). the persistent in vivo metabolite of DDT. Polychlorinated biphenyls (PCBs) and 2,3,7.8-tetrachlorodibenzo-p-dioxin (TCDD) also could disrupt development via several mechanisms. These compounds were selected for review because they are among the most extensively studied HAAs. The work discussed below illustrates the developmental and reproductive effects that can be caused by exposure to these estrogenic, antiestrogenic, and antiandrogenic agents in vivo. The list of compounds discussed in this chapter is by no means complete, and it might not even be representative of all HAAs. The human studies evaluated here involve exposure to DDE, PCBs, and TCDD. Also evaluated are data on regional and temporal variations in sperm concentration in human populations. There is a discussion of adverse reproductive effects observed in wildlife populations. Where available, laboratory studies related to these findings are presented. The section on wildlife describes effectscontinue

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Page 121 of exposure to HAAs in some representative vertebrate species: fish; birds; alligators; turtles; salamanders, frogs, and toads; and the Florida panther. In evaluating the information, it is essential to examine the data that link reproductive and developmental effects to critical periods of exposure and to concentrations ordinarily found in the environment; however, for the most part this information is not available (see Chapter 1 for more details). In the cases where cause-and-effect relationships can be established or posited, known and suspected toxicologic mechanisms are discussed. Laboratory Animal Studies of Selected HAAs Dichlorodiphenyltrichloroethane (DDT) Immature female rats injected intraperitoneally with a single dose of the estrogenic DDT isomer o,p'-DDT at 1 mg/kg had a significant increase in uterine wet weight (Welch et al. 1969), and newborn female rats injected subcutaneously with 1 mg/d o,p'-DDT on d 2-4 after birth had early onset of puberty and accelerated loss of fertility, referred to as ''delayed anovulatory syndrome" (Heinrichs et al. 1971). Gellert et al. (1974) report that subcutaneous injection of 0.1 mg of o,p'-DDT on d 2-4 of life led to marked impairment of fertility and reduced weight of prostate and seminal vesicles in male rats. Male offspring from pregnant rats fed 100 mg/kg/d p.p'-DDE, the antiandrogenic metabolite of DDT. on gestation d 14-18 had reduced anogenital distance and had nipples (androgen normally blocks nipple retention in male rodents) (Kelce et al. 1995). Weanling male rats given daily doses of 100 mg/kg/ d p,p'-DDE by gavage until d 57 had delayed onset of puberty, and castrated adult rats given daily doses of 200 mg/kg/d p,p'-DDE by gavage for 4 d had decreased seminal vesicle and prostate weight (Kelce et al. 1995). Exposure to DDT during gestation also has been shown to impair locomotor ability in mice (Tilson et al. 1979) and learning in rats (Lilienthal et al. 1990; Lilienthal and Winneke 1991) and monkeys (Schantz and Bowman 1989; Schantz et al. 1989). These studies are discussed in Chapter 8. The mechanism by which DDT, or DDE, causes structural or functional abnormalities of the reproductive system in laboratory animals is still poorly understood. As discussed in Chapter 2, the DDT isomer o,p'-DDT has estrogenic properties: the DDT metabolite p,p'-DDE acts as an antiandrogen. Kelce et al. (1995) reported that a concentration of 3.5 µM p,p'-DDE occupied 50% of androgen receptors in rat prostate cells. This was approximately 200 times lower than the concentration of p,p'-DDE required to occupy 50% of estrogen receptors in rat uterine cells. Soto et al. (1997) also showed that p,p'-DDE is a partial agonist of estrogen. They reported that a 10-µM dose of p,p'-DDE was required to produce an increase in proliferation of MCF7 human breast-cancer cells, but the increase in proliferation was only 25% of the maximum proliferative responsecontinue

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Page 122 seen at saturating doses of estradiol. Taken together, these findings show that the capacity for p,p'-DDE to bind to and interfere with the functioning of androgen receptors is considerably greater than its capacity to bind to estrogen receptors and stimulate estrogenic responses. The primary activity of p,p'-DDE as an HAA is thus as an environmental antiandrogen, not as an environmental estrogen. The antiandrogenic properties of p,p'-DDE might be of greater importance than the estrogenic properties of DDT on developing animals. p,p'-DDE persists for decades in tissues, whereas estrogenic o,p'-DDT is much less commonly detected in human serum (Stehr-Green 1989). However, exposure to o,p'-DDT, as well as to other nonpersistent pesticides (methoxychlor, for example), during critical periods in development affects fetal development in mice, and some effects, such as changes in territorial behavior, become apparent only in adulthood (vom Saal et al. 1995). Methoxychlor Methoxychlor is an insecticide used in home gardens and on crops and livestock (ATSDR 1994). The effects of methoxychlor on the reproductive systems of female rats and mice have been studied extensively (Cummings 1997). Methoxychlor causes adverse effects on fertility, early pregnancy, and in utero development. Accelerated pubertal ovulation, persistent vaginal cornification. accelerated loss of fertility, and abnormal cell types in the uterus and oviducts have been observed after neonatal administration of doses as low as 0.5 µg/d per neonate (Welch et al. 1969; Gellert et al. 1974: Gray et al. 1989; Eroschenko and Cooke 1990; Gray 1992). When administered to mated female rats during the peri-implantation period, a 300-mg/kg/d (approximately 100 mg/d) dose of methoxychlor completely blocked implantation of embryos (Cummings and Gray 1989; Gray et al. 1989; Cummings 1990). Exposure to methoxychlor during development also leads to changes in the reproductive system and behavior of male rats and mice. Male offspring of mice fed 20 µg/kg/d methoxychlor in oil during the last third of pregnancy exhibited an increase in territorial marking behavior in adulthood, similar to the effect observed with a 20-µg/kg/d dose of o,p'-DDT and a 20-ng/kg/d dose of DES (vom Saal et al. 1995). Daily intraperitoneal injection of 1 mg of methoxychlor to male mice during the first week after birth led to reduced serum testosterone concentrations and to reduced DNA content in prostate and seminal vesicles in adulthood (Cooke and Eroschenko 1990). Administration of 50 mg/kg/d methoxychlor to female rats throughout pregnancy and lactation resulted in smaller testes, epididymides, and reduced sperm count in male offspring (Gray et al. 1989; Gray 1992). The mechanism by which methoxychlor affects the reproductive system and reproductive behavior of laboratory animals is not understood. Methoxychlor has estrogenic effects in vivo only after demethylation in the liver to mono-soft

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Page 123 hydroxy-methoxychlor (30% of administered dose) or bis-hydroxy-methoxychlor (23% of administered dose) (Kapoor et al. 1970); the more potent estrogenic metabolite is bis-hydroxy-methoxychlor (Welch et al. 1969; Bitman and Cecil 1970; Bulger et al. 1978; Bulger and Kupfer 1983). Unlike p,p'-DDE, methoxychlor is not persistent in vivo (most is cleared within 24 h), although it is relatively persistent (a number of months) in soil (Muir and Yarechewski 1984; ATSDR 1994). Chlordecone Chlordecone [decachlorooctahydro-1,3,4-metheno-2H-cyclobuta(cd)pentalen-2-one] was used in the 1960s and 1970s to control insect pests of bananas, citrus trees without fruit, tobacco, and ornamental shrubs (ATSDR 1995). The insecticide mirex, which is similar in structure to chlordecone, has been used in much greater quantities to combat fire ants. When a 15-mg/kg/d dose of chlordecone was fed to pregnant rats on d 14-20 of gestation, 12 of 21 female offspring developed persistent vaginal estrus; the other nine rats were anovulatory at 6 mo of age (Gellert and Wilson 1979). The effects were consistent with those seen after exposure to estrogen, but no estrogenic effects were observed among male offspring. In a study of postnatal exposure, constant vaginal estrus was induced in mature female rats fed 1.5 mg/ kg/d chlordecone for 7 d (Hammond et al. 1979). Mature female mice injected with 125 µg/d chlordecone on postnatal d 1-10 developed complete cornification of the vaginal epithelium—this was similar to the effects caused by treatment with 10 µg/d estradiol (Eroschenko and Palmiter 1980). In males, spermatogenesis was completely suppressed by estradiol and was reduced by chlordecone. Injection of either 0.2 or 1 mg/kg/d chlordecone into newborn female rats on d 2 and 3 of life led to early onset of puberty and accelerated loss of cyclicity (Gellert 1978a). In a study using quail, Eroschenko (1981) fed male quail a diet containing 200 ppm chlordecone for 3 wk and reported a significant increase in testes weight, due to edema, with dilation of the seminiferous tubules and erosion of the germinal epithelium. Abnormal sperm was also observed. With exposure for 6 wk, the testes in some animals began to atrophy. The binding affinity of chlordecone to the estrogen receptor is approximately 0.02% (5,000-fold) lower relative to estradiol (Eroschenko and Palmiter 1980). Vinclozolin Vinclozolin (3-(3,5-dichlorophenyl)-5-methyl-5-vinyl oxazolidine-2,4-dione) is a dicarboximide fungicide widely used to combat damage to a variety of commodities, such as fruits, vegetables, hops, and turf. When vinclozolin was administered via gavage to pregnant and lactating rats from gestation d 14 throughcontinue

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Page 124 postnatal d 3 at 100 or 200 mg/kg/d, male offspring were indistinguishable from female offspring on external examination at birth, and the males retained nipples (Gray et al. 1994). These findings indicate that the masculinizing effects of androgen on the external genitalia and the defeminizing action of androgen on the development of nipples were blocked by vinclozolin. In adulthood, due to gross abnormalities of the internal and external genitalia, the vinclozolin-treated males were infertile. Vinclozolin is an androgen receptor antagonist (Kelce et al. 1994). After ingestion, it is degraded to metabolites that compete with endogenous androgen for binding to androgen receptors. Polychlorinated Biphenyls (PCBs) PCBs are no longer legally manufactured in the United States, but large quantities were produced for use in such products as electrical transformers and capacitors. Some of the more highly chlorinated of the 209 potential PCB congeners are highly persistent, and they bioaccumulate in the food chain. Rats exposed to PCB mixtures early in life can develop reproductive effects similar to those caused by DES exposure (Bitman and Cecil 1970: Sager 1983: Sager et al. 1987, Subramanian et al. 1987; Lundkvist 1990: Jansen et al. 1993: Bergeron et al. 1994: Birnbaum 1994: Gray et al. 1995; Li and Hansen 1996). Male offspring of rats fed 32 or 64 mg/kg/d PCBs during lactation had significantly reduced seminal vesicle weight and significantly larger testes (Sager 1983; Sager et al. 1987). Those effects might be due to changes in thyroid hormone levels (Jannini et al. 1993). When treated males were mated with unexposed females, there was a significantly lower incidence of implantation, a significantly lower number of live births, and a significantly greater rate of resorption. Female offspring of dams exposed to 32 or 64 mg/kg/d PCBs during lactation had delays in puberty, vaginal opening, and first estrus (Sager and Girard 1994). At maturity, uterine wet weight was reduced at all stages of the estrous cycle, and fertility was impaired because of reduced success at the pre- or postimplantation stage. Female rats treated intrapcritoneally with PCBs also had significant increases in uterine weight (Jansen et al. 1993). When female guinea pigs were force-fed 2.2 mg/d ( 1.8-3.2 mg/kg/d) PCBs during gestation, female offspring had delayed vaginal opening and male offspring had significantly reduced absolute and relative testis weight (Lundkvist 1990). Pre- and postnatal exposure to coplanar PCBs can modulate thyroid hormone concentrations and uptake. When pregnant rats were orally administered 0.2, 0.6. or 1.8 mg/kg/d HCB (3,3',4,4',5.5'-hexachlo-obiphenyl) or a combination of 1 mg/kg/d TCB (3,3',4.4'-tetrachlorobiphenyl) and 0.6 mg/kg/d HCB, there were decreases in fetal, neonatal, and weanling plasma total thyroxine (T4) and free T, concentrations, which indicated an increase in peripheral T4 metabolism (Morsecontinue

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Page 125 et al. 1993). An increase in the activity of type II thyroxine 5'-deiodTMase in brain homogenates also was observed, which suggests that local hypothyroidism occurs in the brains of fetal and neonatal rats exposed to these PCBs. The PCB mixture Aroclor 1254, administered orally at doses of 5 or 25 mg/kg/d to pregnant rats on d 10-16 of gestation, caused the selective accumulation of a hydroxylated PCB metabolite (2,3,3',4,5'-pentachloro-4-biphenylol) in fetal plasma and brain; this was believed to be the cause of the reductions of fetal plasma and brain T4 concentrations (Morse et al. 1996). In a reproductive toxicity study of rhesus monkeys, 80 females were fed 5-80 µg/kg/d Aroclor 1254 prior to breeding, during breeding (with untreated males), and after breeding, for 6 yr. There was a significant dose-related decreasing trend in conception rate and a significant dose-related increasing trend in fetal mortality (Arnold et al. 1995). Maternal age was not a confounding factor. It was noted during this study that while some of these animals had endometriosis the incidence or severity of the lesions could not be related to the PCB treatment. Similarly, a recent study in women led to the conclusion that exposure to PCBs and to chlorinated pesticides is not associated with endometriosis in the general population (Lebel et al. 1998). In addition to reproductive effects, prenatal exposure to PCBs has been shown to cause deficits in neurodevelopment, such as impaired learning and altered activity levels in rats fed PCB-contaminated fish (Tilson et al. 1990; Daly 1992). These studies are described in greater detail in Chapter 6. As discussed in Chapter 2, there is evidence that PCB mixtures and congeners and hydroxy-PCBs can have estrogenic and antiestrogenic properties. In addition, coplanar PCBs are antiestrogenic through aryl hydrocarbon-estrogen receptor crosstalk (Krishnan and Safe 1993). However, the significance of results from laboratory studies of PCBs are difficult to interpret because most PCB extracts from environmental samples do not resemble commercial PCB mixtures used in laboratory studies (WHO 1993). As PCBs are cycled through the environment, the various congeners are gradually redistributed. The most chlorinated congeners, which are typically the most toxic, accumulate preferentially. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) TCDD is a byproduct of the production of chlorinated products such as herbicides and wood preservatives, the incineration of trash containing papers and plastics, and the burning of fossil fuels (e.g., IARC 1997). A series of studies (Mably et al. 1992a,b,c) has shown effects on male rats whose mothers were given single oral doses of TCDD ranging from 0.064 µg/kg to 1 µg/kg on gestation d 15. At doses as low as 0.16 µg/kg, impaired sexual differentiation was observed in male fetuses, including a decrease in circulating testosterone and in anogenital distance at birth (Mably et al. 1992a). Effects on sexual behavior in male offspring in adulthood included changes in mounting, intromitting, numbercontinue

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Page 126 of ejaculations, and latency to ejaculation, as well as in the exhibition of the female sexually receptive posture (lordosis) (Mably et al. 1992b). In addition, a dose-related decrease in the weight of the testis and epididymis was observed. There was also a decrease in daily sperm production in male offspring of pregnant rats receiving a single dose of 1 µg/kg, although there was no effect on fertility (Mably et al. 1992c). Maternal exposure to a single oral dose of 1 µg/kg TCDD on gestation d 8 or 15 caused delayed puberty, partial clefting of the penis, and "thread" tissue development across the opening of the vagina in female offspring of rats (Gray and Ostby 1997). Ovarian weight was significantly reduced. In male offspring, the same treatment on gestation d 15 caused delayed puberty and reductions in ejaculated and epididymal sperm counts and in sex accessory gland size (Gray et al. 1995). Chronic oral exposure to TCDD caused endometriosis in rhesus monkeys, with incidence and severity related to dose (Rier et al. 1993). Specifically, adult female monkeys were administered 2.5 x 10-7 and 1.25 x 10)-6 mg/kg/d in their food over 4 yr. Ten years later, moderate to severe endometriosis was found during laparoscopic examination in three of seven monkeys exposed to 5 ppt TCDD and in five of seven monkeys exposed to 25 ppt TCDD. None of the control monkeys showed severe disease. The mechanism by which TCDD causes reproductive impairments and affects reproductive behavior is poorly understood. Because outcomes of prenatal TCDD exposure in rats appear similar to effects seen after treatment with DES (Peterson et al. 1993; Eskenazi and Kimmel 1995; Gray et al. 1995), the general assumption that TCDD acts as an estrogen antagonist might not apply to all effects of TCDD. In contrast, studies have shown that TCDD causes antiestrogenic responses in the rodent uterus (Gallo et al. 1986; Romkes et al. 1987: Umbreit et al. 1988, 1989; Astroff and Safe 1990; DeVito et al. 1992) and in breast cancer cells (Biegel and Safe 1990; Harris et al. 1990: Safe et al. 1991: Gierthy et al. 1993: Merchant et al. 1993) and inhibits mammary tumor growth in rodents (Kociba et al. 1978; Gierthy et al. 1993; Holcomb and Safe 1994; Tritscher et al. 1995). TCDD causes antiestrogenic effects via the aryl hydrocarbon (Ah) receptor signaling pathway, which has been characterized at the molecular level (Krishnan et al. 1995; Gillesby et al. 1997). Bisphenol A Bisphenol A (4,4'-(1 -methylethylidene)bisphenol) is a monomer used in the manufacture of polycarbonate plastic, resins, and some dental sealants; it is an additive in numerous other products. In a developmental toxicity study with pregnant rats and mice (Morrissey et al. 1987), gastric intubation of 1,250 mg/kg/ d bisphenol A on d 6-15 of gestation significantly decreased maternal body weight, increased maternal mortality and resorption of fetuses, and decreased thecontinue

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Page 127 weight of surviving pups in mice but not in rats. There were no observable malformations in exposed fetuses. In a subsequent study that used much lower doses of bisphenol A (Nagel et al. 1997), male offspring of pregnant mice fed 2 µg/kg/d bisphenol A on d 1 11-17 of gestation had significantly increased prostate weight as adults. In addition, 2 µg/kg/d bisphenol A produced significant enlargement of the preputial glands in male offspring, whereas the epididymides were significantly reduced. A dose of 20 µg/kg/d bisphenol A reduced daily sperm production per gram of testis by 20%, while daily sperm production uncorrected for testis weight was not significantly different (vom Saal et al. 1998). The investigators suggested that these doses could be within the range encountered by humans, as evidenced by amounts detected in the saliva of 18 patients (amounts ranged from 90 to 931 µg) treated with dental sealants (Olea et al. 1996) and in a few lacquer-coated cans of vegetables (amounts ranged from 0 to 23 µg/can) (Brotons et al. 1995). Bisphenol A does not bind to plasma-binding proteins or other components of blood to the same degree that estradiol does, and therefore, it passes more readily from blood into cells (Nagel et al. 1998). The inhibition of the uptake of steroids into cells by components of blood is particularly important during fetal life in rats, when the concentrations of gonadal and adrenal steroids are great but the bioactive fraction of steroid that can enter cells is maintained at a low concentration (vom Saal et al. 1992). Effects of bisphenol A are seen in vitro in human breast-cancer MCF7 cells at 10-8 M (10 nM) or 2.3 ppb (molecular weight, 228) (Krishnan et al. 1993; Olea et al. 1996), and in F344 rat prolactin-secreting GH3 pituitary cells at 1 nM (Steinmetz et al. 1997). The finding of increased prostate size in male offspring of pregnant female mice fed 2µg/kg of bisphenol A (Nagel et al. 1997) is similar to that found in tests with DES at a dose of 0.02 µg/kg (vom Saal et al. 1997), which shows that bisphenol A is approximately 100-times less potent than DES when fed to pregnant mice. The same 2 µg/kg dose of bisphenol A fed to pregnant female mice also advanced the timing of puberty in female offspring, an effect seen with higher doses of other estrogenic chemicals (Howdeshell et al. 1999). In another study (Steinmetz et al. 1997), estradiol and bisphenol A were administered to F344 and Sprague-Dawley rats via subcutaneous Silastic capsules, and serum prolactin, which is elevated by estrogen treatment, was measured (along with numerous other responses). Bisphenol A induced hyperprolactinemia in F344 rats but not in Sprague-Dawley rats. In the F344 rat, estradiol increased serum prolactin concentrations 10-fold and bisphenol A increased it 7- to 8-fold over controls. These findings are consistent with the findings of vom Saal et al. (1997) comparing bisphenol A and DES in fetal mice. Those findings suggest that bisphenol A is bioactive within the range of human exposure (Walent and Gorski 1990; Krishnan et al. 1993; Brotons et al. 1995: Olea et al. 1996; Takao et al. 1999).break

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Page 128 Recently, Cagen et al. (1999) conducted a study to evaluate the effects of low doses of bisphenol A on sexual development in male mice. To the extent possible, the study protocol duplicated the studies of Nagel et al. (1997) and vom Saal et al. (1998) for all factors indicated as critical by those investigators. Some differences between the two studies include the source of the CF,, mice, the number of doses of bisphenol A, the methods used to determine sperm count, the age of mice at necropsy, and the way the animals were housed. An additional positive control group of mice was dosed with 0.2 µg/kg/d DES. No effects on testes histopathology, daily sperm production or sperm-production efficiency (i.e., daily sperm production per gram of testis), or on prostate, preputial, seminal vesicle, or epididymis weights were observed in the bisphenol A treated groups. In addition, no adverse effects were observed in any of these parameters in the group tested with DES. Thus, this study failed to replicate the results of vom Saal et al. (1997, 1998) and Nagel et al. (1997). The reason for the discrepancies in the findings is a subject of controversy and cannot be resolved at this time. Octylphenol Octylphenol (p-(1,1,3,3-tetramethylbutyl)phenol) is an alkylphenol used in its ethoxylated form (octylphenol ethoxylate) in a variety of products such as detergents and plastics. In one study (Sharpe et al. 1995), female rats were given drinking water containing 100 or 1,000 µg/L octylphenol before pregnancy, during gestation, and throughout lactation, and the effects on testicular size and spermatogenesis in male offspring in adulthood were investigated. Intake of octylphenol based upon water intake was calculated only for the group exposed to 1,000 µg/L and was reported to range from 129 µg/kg/d in the first 2 d after birth to 367 µg/kg/d just before weaning. There were significant decreases in absolute testis weight, in the ratio of testis-to-kidney size, in relative ventral prostate weight, and in daily sperm production in male offspring exposed to 1,000 µg/L octylphenol, and there was a significant reduction in relative prostate weight at both concentrations. When male offspring were treated postnatally only with 1,000 µg/L octylphenol on d 1-22 after birth, there was a significant reduction in average and relative testis weight (Sharpe et al. 1995). More recently, the authors reported their inability to replicate their findings (Sharpe et al. 1998). While they expressed continued confidence in their original publication, they hypothesized that the inability to replicate the work may have been due to changed biologic factors of which they were unaware and unable to control (Sharpe et al. 1998). In another study, there was no effect on prostate weight of male offspring of pregnant mice fed 2 and 20 µg/kg/d octylphenol on d 11-17 of gestation (Nagel et al. 1997), although there was a significant decrease in daily sperm production (vom Saal et al. 1998).break

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Page 129 Butyl Benzyl Phthalate (BBP) BBP is used as a plasticizer for cellulose resins, polyvinyl acetates, polyurethanes, and polysulfides and in regenerated cellulose films for packaging. When 1,000 µg/L BBP was administered in drinking water to female rats before mating and throughout lactation (nominal intake based on water intake ranged from 126 µg/kg/d in the first 2 d after birth to 366 µg/kg/d just before weaning), there was a small but significant reduction in mean testicular size and a reduction in daily sperm production in male offspring at d 90-95 (Sharpe et al. 1995). Similar findings were reported in rats treated with DES (100 µg/L in drinking water; intake was not determined), which was evaluated in the same study. The researchers note that, although these estrogenic chemicals exert similar effects on testis size and daily sperm production, there is no evidence that the effects are caused by the compounds' estrogenicity. In a subsequent study, Ashby et al. (1997) failed to find any effects of BBP, despite very similar or identical protocols. Female rats were administered BBP (average intake 182.6 µg/kg/d) in drinking water during gestation and lactation, and their offspring were monitored for 90 d. DES (8.6 µg/kg/d) affected the sexual development of male and female pups, causing changes in anogenital distance; average day of vaginal opening and prepuce separation; weight of the uterus, testis, and accessory sex glands; and caudal epididymis sperm count and homogenization-resistant testicular sperm count. BBP had no effect other than to cause a slight advance in the average day of vaginal opening and a small increase in male anogenital distance, but those effects were attributed to the increased weight of the pups treated with BBP. The reason for the discrepancy in findings between the study by Sharpe et al. (1995) and that of Ashby et al. (1997) is unknown. Di-n-Butyl Phthalate Di-n-butyl phthalate (DBP) has been characterized as a reproductive and developmental toxicant in several studies (e.g., Cater et al. 1977; Ema et al. 1994, 1995). causing fetal death and skeletal malformations (predominantly cleft palate) in rats. In addition, DBP has been shown to cause testicular atrophy, early sloughing of germ cells, and vacuolization of Sertoli cell cytoplasm (Cater et al. 1977; Fukuoka et al. 1989). Immature rats appear to be more susceptible to these effects than adult rats (Gray and Gangolli 1986; Creasy et al. 1987). Recent studies have investigated the effects in more depth. DBP was tested in the National Toxicology Program's Reproductive Assessment by Continuous Breeding protocol using Sprague-Dawley rats (Wine et al. 1997). DBP was administered in the diet to male and female rats continuously. with an average daily intake of 52, 256, and 509 mg/kg for males and 80, 385. and 794 mg/kg for females, respectively. Breeding pairs (F0 generation) were matedcontinue

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Page 160 Birds Studies examining the birds of the Laurentian Great Lakes provided much of the initial data that led to the concept of hormonally active environmental pollutants (Colborn 1990). Contaminant-related effects on the reproductive health of several avian species appear to involve modifications during embryonic development (Colborn et al. 1993; Fox 1992). The studies of interest have shown alterations in sexual behavior, abnormal reproductive morphology, severe developmental abnormalities associated with growth and metabolism, and eggshell thinning. An important aspect of understanding the potential effects of exposure to HAAs on the development of avian reproductive systems is that estrogen plays an important role in regulating the course of development and adult functioning of the gonads and accessory reproductive structures (Fry 1995). Estradiol in conjunction with other endocrine and paracrine factors, is implicated in the unilateral development of the left ovary and regression of the right ovary. Estradiol also influences whether the embryonic tissues that differentiate into oviducts and shell glands persist or regress, via the interaction of estradiol with Müllerian inhibiting hormone; the action of Müllerian inhibiting hormone is inhibited by estradiol. The active differentiation of female reproductive structures under the action of estrogen is in direct contrast to mammals, in which males are the heterogametic and differentiating sex, and estradiol is not required for differentiation of the ovary, although in mammals, estrogen and functioning estrogen receptors are required for subsequent normal functioning of the ovary (Lubahn et al. 1993). Thus, embryonic exposure to environmental estrogens or other HAAs can generate end points in birds that are different from what might be expected in mammals or other vertebrate species that do not rely on estrogen-induced sexual differentiation. This concept of varying end points among species is essential to explaining the effects observed in wildlife. Such variations do not preclude the use of data from birds or other wildlife to predict abnormalities or effects in other species. The information can be used to help determine which end points are the most likely for a given species and to explain why effects are seen in some groups and not others. Thus, an understanding of the role of estrogen in avian development leads to the prediction that exposure to environmental estrogens could alter sex ratios and feminize male birds even though different outcomes would be predicted in other vertebrates. Sex Ratios During the 1950s and 1960s, when organochlorine (e.g., o,p'-DDT, PCBs, dioxin) contamination was at its greatest in the United States, an increased incidence of female-female pairings in gull populations was observed in colonies in California and the Great Lakes (Schreiber 1970; Harper 1971; Hunt and Huntcontinue

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Page 161 1973; Gress 1974: Hand 1980; Shugart 1980; Fitch and Shugart 1983; Fry et al. 1987). This phenomenon is usually estimated by documenting the number of nests that contain five or more eggs (supernormal clutch); a single female gull typically lays one to three eggs. The most dramatic and well-documented example of sex skew occurred in the western gull population on Santa Barbara Island in California from 1968 to 1978 (Hunt et al. 1980). The adult sex ratio in that population was measured by laparotomy of 856 captured birds to be 0.26 males to females. The investigators also calculated the male-to-female ratio by estimating the number of nests on the island (896), the number of nonbreeding birds (200), and, based upon the number of nests with more than three eggs, the percentage of female-female pairs (15%). Using those estimates, the male-to-female ratio was 0.67. Because many birds laid fewer than normal numbers of eggs in 1978, the investigators believe that the estimate for female-female pairs might been too low. Therefore, the ratio of males to females was probably between 0.26 and 0.67. A supernormal clutch incidence of 0.6% to 1% was documented in northeastern Lake Michigan herring gulls from 1978 to 1981 (Shugart 1980: Fitch and Shugart 1983). Both the California population and the Great Lakes gulls were exposed to great levels of organochlorine contamination, including DDT, during the 1950s to the 1970s (Fry and Toone 1981). Several historical studies have been done to investigate the occurrence of supernormal clutches in gulls using literature sources and museum specimens to determine whether incidences have actually changed in the pre- and post-DDT era. The incidence of supernormal clutches has decreased significantly for many species of terns throughout the United States (Conover 1984a). Supernormal-clutch incidence had only increased significantly in western gulls, herring gulls nesting in the Great Lakes, and Caspian terns breeding in the United States since 1950. Supernormal clutches were a regular occurrence in ring-billed and California gulls before the DDT era, and their occurrence has not changed over time (Conover and Hunt 1984a). In contrast, supernormal clutches were not found regularly in western or herring gulls until after 1950, and the sex ratio for their populations as a whole has changed dramatically for both species toward an excess of females since then. Sex skew and its effects on the population dynamics of gulls are discussed in detail in Chapter 10. There are a number of hypotheses concerning sex skew. In general, female-female associations in gull colonies are believed to occur when there is a relative shortage of breeding male gulls available. Experimental manipulation of sex ratios in gull colonies by selectively removing males from stable colonies has demonstrated that sex ratio skew alone is sufficient to cause a proportion of the excess females to pair (Conover and Hunt 1984b). Sex skew toward females in western and herring gulls could be due to a differential mortality between males and females; however, such a differential mortality has not been well documented. It is possible that male gulls could be more susceptible to poisoningcontinue

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Page 162 from persistent organochlorine contaminants. Male western gulls weigh about 25% more than females on average, and they feed higher up on the food chain (Pierotti 1981). Also, male gulls do not have the ability to excrete lipophilic contaminants by egg-laying. For these reasons, it is expected that male gulls might accumulate greater body burdens of toxicants throughout their lifetimes than females gulls do. It has also been suggested that the skewed sex ratios observed in western gulls in California and in Great Lakes herring gulls might have been caused by estrogenic contaminants, such as DDT, in the environment, either due to a differential male mortality or a feminization of male embryos which resulted in chemical sterilization and a failed recruitment into the breeding population (Fry et al. 1987). This is a plausible hypothesis, but there is no direct evidence to support it. Another hypothesis is that some females might have paired with the wrong sex due to chemically induced masculinization. However, a behavioral study of western gulls in Santa Barbara did not find significant differences in behavior between females mated with other females and those paired with males (Hunt et al. 1984). In conclusion, there is good evidence to suggest that there has been a fundamental change in the sex ratio of several North American gull populations in the post-DDT era, such that there is an overabundance of females in some breeding colonies. The observations that the colonies most affected were in areas of great DDT contamination and that a few DDT congeners have produced abnormal gonadal development in laboratory studies support the hypothesis that environmental contaminants may have played a role in the sex ratio skew. Alterations in Behavior Behavioral abnormalities observed in the wild include aberrant parental behaviors, such as less inclination to sit on eggs or to defend nests, which was observed in herring gulls in Lake Ontario (Fox et al. 1978). Those alterations were sufficient to account for the high incidence of egg loss observed in this population. Because high levels of chemical contamination were found in the gulls, it was suggested that HAAs might be responsible for the behavioral alterations. Laboratory experiments with birds exposed to hormones and some environmental pollutants suggest that this hypothesis is plausible. For example, male Japanese quail embryos injected with 1 µg estradiol or 500 µg testosterone were completely demasculinized and were behaviorally indistinguishable from females (Adkins 1979; Adkins-Regan 1987). As adults, they failed to mount, crow, or strut. This effect occurred only if treatment was given before d 12 of the 18-d incubation period (Adkins 1979), and it is believed to result from a fundamental change in the neural substrate underlying behavior that confers a differential responsiveness to the activating effects of testosterone in adulthood. Testosterone treatment restores copulation in castrated adult males but is without effect in females. Female quail treated with an antiestrogen beforecontinue

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Page 163 hatching can be masculinized and as adults will mount other females (Adkins-Regan 1987). Sexual differentiation of behavior has been extensively studied in the zebra finch, which exhibits sexual behavioral dimorphism; that is, normally only males sing, dance, and mount. The brain of this finch is sexually dimorphic. The telencephalic nuclei greater vocal center, nucleus robustus archistriatalis, nucleus magnocellularis of the anterior neostriatum, and area X of the lobus paraolfactorius are larger and more extensively connected in males than in females, and are essential for learning and production of the complex vocalizations of this species (Simpson and Vicario 1991a). The administration of estradiol to female zebra finches during the first week after hatching results in a profound organizational masculinization of brain and behavior (Gurney and Konishi 1980; Simpson and Vicario 1991 a; Adkins-Regan et al. 1994), including neural masculinization of telencephalic nuclei that sets up a functional circuit in females similar to that in males, which enables them to learn and produce complex vocalizations (Simpson and Vicario 1991b). When the treated females are stimulated as adults with testosterone, they engage in male behavior of singing and dancing (Adkins-Regan et al. 1994). Males treated with estradiol during the first week after hatching are demasculinized, and they fail to mount as adults (Adkins-Regan et al. 1994). Thus, the pattern of sexual-behavioral differentiation in the zebra finch is quite complex. It is clear from these studies and those involving quail that the process of sexual behavioral differentiation in birds is sensitive to exogenous hormones, and that hormonal manipulation can result in a profound and permanent change in reproductive behavior in both sexes. In laboratory studies with contaminants, ring doves were fed mixtures of DDE, PCBs, mirex, and photomirex (contaminants found in salmon and gulls of Lake Ontario) during mating. The feed of the low-dose group contained 8 ppm Aroclor 1254, 1.67 ppm DDE, 0.297 ppm mirex, and 0.0954 ppm photomirex; the high-dose diet contained 29.03 ppm Aroclor 1254, 4.61 ppm DDE, 0.897 ppm mirex, and 0.324 ppm photomirex. The doves had reduced or delayed behaviorally induced increases of sex hormones, females failed to respond normally to male courtship behavior, pairs spent less time building their nests, and pairs receiving the greatest dosage spent less time feeding their young (McArthur et al. 1983). There was a marked dose-related decrease in fledgling success, and the breeding cycle was greatly asynchronous. In other studies, adult breeder doves fed PCBs exhibited aberrant incubation (Peakall and Peakall 1973) and courtship (Tori and Peterle 1983). Thus, there is evidence from laboratory studies that environmental contaminants in the Great Lakes region could cause behavioral anomalies in breeding synchrony, nest construction, incubation attentiveness, and parental care at ambient concentrations but these effects are not necessarily attributable to their hormonal activities.break

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Page 164 Abnormal Reproductive Morphology Most studies of abnormal breeding in gull populations were conducted during the mid-1970s. Fifty-seven percent of the male gull embryos collected from Scotch Bonnet Island, Canada, in 1975 and 1976 had testicular feminization (Fox 1992). Eggs at that site were contaminated with dioxins, PCBs, and mirex (Gilman et al. 1979; Fox 1992). One study of a tern colony also showed a high incidence of abnormal tests indicative of estrogenic exposure in ovo (Calambokidis et al. 1985; Nisbet et al. 1996), but the contaminants that could mediate these effects have not been identified. However, the significance of these findings is unclear because there are reports (going back to the 1800s) of apparently abnormal testes in terms as embryos or hatchlings, and this might be a normal condition that disappears as the birds age (Hart 1998). Some studies have evaluated the reproductive morphology of adult birds. For example, 31 adult female glaucous-winged gulls collected in 1984 from Tacoma, Washington, adjacent to the Commencement Bay, Puget Sound (a PCB-and heavy-metal-contaminated Superfund site), were trapped on their nests and killed for gonadal inspection (Calambokidis et al. 1985). The right oviducts of these gulls were found to be persistent and large. The length of the right oviduct was correlated with the estimated chemical contamination (Fry et al. 1987). However, the significance of these data are unclear, as all birds were successfully incubating clutches. Furthermore, the most severe category of oviduct enlargement was rated as greater than 10 mm long; the literature indicates that a vestigial right oviduct of 9-10 mm is normal in the herring gull (Boss and Witschi 1947). An attempt to correlate alterations in testes in male birds with organochlorine contamination in this gull population was inconclusive (Fry et al. 1987). Ovotestis formation in male embryos and retention of the right oviduct in female embryos also were observed in experimental studies of gull eggs injected with hormones such as estradiol (Fry and Toone 198 1) and DES (Boss and Witschi 1947) and with environmental contaminants such as methoxychlor and DDT (Fry and Toone 1981). Chicken and quail eggs injected with DDT showed similar effects (Lutz-Ostertag and David 1973). Because concentrations of DDT (2-100 ppm) found in the eggs of wild gulls caused effects consistent with those induced by estradiol and DES (Fry and Toone 1981), it is plausible that DDT or other estrogenic contaminants could be responsible for the effects observed in the wild. Most morphologic abnormalities in the wild have been found near areas identified as hot spots of organochlorine contamination. Residues of PCBs, TCDD-EQ, and DDT are approximately 10-fold greater than those in other locations (Giesy et al. 1994b). Concentrations of many of the residues are declining in the Great Lakes, but are still among the highest. Green Bay and Saginaw Bay also have hot spots, where concentrations of organohalogen compounds are significantly greater than the Great Lakes as a whole (Giesy et al. 1994b).break

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Page 165 Growth and Development Abnormalities A group of embryonic abnormalities directly related to contaminant exposure in some fish-eating birds has been defined as a specific syndrome. GLEMEDS (Great Lakes embryo mortality, edema, and deformity syndrome) (Gilbertson and Fox 1977; Gilbertson et al. 1991). GLEMEDS involves a consistent pattern of subcutaneous edema, beak malformations, cardiac edema, and skeletal malformations. The expression of this syndrome in bald eagle, cormorant, gull, and tern chicks is correlated with dioxin toxic equivalents of some PCB congeners that are primarily the result of maternal bioaccumulation from eating contaminated fish and resultant deposition of coplanar PCB congeners in the eggs (Gilbertson et al. 1991). Adults from populations in which chicks have GLEMEDS have shown abnormal plasma thyroid hormone concentrations and thyroid morphology (Fox 1992), but no relationship between thyroid hormonal disfunction and GLEMEDS has been found. Several sources of organochlorines have been controlled in response to regulatory action in the early 1970s, and concentrations of DDT and PCBs in fish tissues decreased approximately 20-fold in the late 1970s and early 1980s. However, no change was found between 1985 and 1992 in chinook salmon (Miller 1994), and these contaminants continue to persist in tissues and the environment today. Eggshell Thinning During the 1960s and 1970s, when the pesticide DDT and its metabolite DDE were present at higher concentrations than today in North America, it was observed that populations of several bird species declined when individuals were unable to successfully incubate eggs because of abnormally thin eggshells (Cooke 1973). Many of these species, such as the double-crested cormorant, have experienced dramatic population increases since DDT was banned from use in the United States (Ludwig 1984; Weseloh and Ewins 1994). It is now well established that the DDT metabolite, DDE, and to a lesser extent other organochlorines, causes eggshell thinning (for a review, see Cooke 1973). Research into the mechanism of DDE-induced eggshell thinning has been extensive (Gould 1972; Peakall et al. 1975; Cooke et al. 1976; Miller et al. 1976; Eastin and Spaziani 1978; Cooke 1979; Lundholm 1980, 1982, 1984a,b,c, 1985, 1987, 1988, 1993, 1994; Lundholm and Mathson 1983; Lundholm and Bartonek 1991, 1992; Haynes and Murad 1985). Some of the postulated mechanisms include premature termination of shell formation, premature oviposition, effects on the protein matrix of the shell, effects on initiation sites of shell formation, enhancement of shellgrowth inhibitors, decrease in carbonate availability for shell formation, effects on progesterone binding in the shell-gland mucosa, and alteration in calcium metabolism of the shell gland. The current hypothesis regarding the mechanism of DDE-induced eggshell thinning is an inhibition of prostaglandins by the shell-soft

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Page 166 gland mucosa (Lundholm and Bartonek 1992). Many of the biochemical end points described above are interrelated, and it has been difficult to determine which end points are the direct targets of DDE and which are merely coinfluenced by its action. However, it does not appear that eggshell thinning is a result of DDE acting as a hormone-receptor agonist or antagonist. The situation is complicated further because sensitivities to DDE-induced eggshell thinning vary among avian species, suggesting that different mechanisms cause eggshell thinning in different species. Florida Panther The possibility that exposure to HAAs affects reproduction in the endangered Florida panther has generated considerable interest. The population size is estimated at 30-50 animals of two genetic strains (O'Brien et al. 1990). Most of the panthers exhibit developmental abnormalities (including congenital heart defects) and defects of the reproductive system (cryptorchidism, low sperm density, and sperm defects) (Barone et al. 1994). Reproductive abnormalities had been attributed to genetic inbreeding (Miththapala et al. 1991; Roelke et al. 1993), but a study by Facemire et al. (1995) examining contaminant loading in female panthers has led investigators to conclude that persistent, bioaccumulated contaminants, such as organochlorines, also could contribute to the problems observed. Because the Florida panther is an endangered species, tissue samples for analysis are rare. Three females were examined after death for concentrations of mercury and several bioaccumulated organochlorine compounds. Concentration ranges of various contaminants found in the muscle (µg/g lipid fresh weight) of the animals were 5.45-57.65 µg/g p,p'-DDE; 7.32- 27.06 µg/g Aroclor-1254; <0.0098-2.00 µg/g oxychlordane; and <0.0098-4.82 µg/g trans-nonachlor. The most frequent developmental abnormality in the Florida panther population is cryptorchidism or testicular nondescent. Cryptorchidism has increased exponentially in male cubs since 1975 (Roelke 1990), and 70% of wild Florida panthers are at least unilaterally cryptorchid, as compared with 20% in the mid-1980s (Roelke et al. 1993). Most of the male panthers exhibiting cryptorchidism have the testis in an inguinal location. A comparative study of free-ranging and captive panthers from Florida, Texas, Colorado, Latin America, and North America indicates that the incidence of cryptorchidism in the Florida panther is more than ten times that found in other populations (Barone et al. 1994). Only two cases of cryptorchidism have been reported in captive populations of North American panthers; the condition has never been reported in any other large felid (Roelke et al. 1993). Cryptorchidism is a heritable trait in some inbred domesticated species (McPhee and Buckley 1934; Claxton and Yeates 1972) that could be a response to an abnormal hormonal environment during embryonic development (Hezmall and Lipshultz 1982; Sharpe and Skakkebaek 1993: Hutson et al. 1994).break

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Page 167 Müllerian-inhibiting hormone has been implicated in testicular descent (Hutson et al. 1994), but no data are available on the influence of HAAs on the synthesis of that hormone during gonadal development. In studies with rats, prenatal exposure to the antiandrogenic metabolites of the fungicide vinclozolin caused hypospadias, cleft penis, and suprainguinal (cryptorchid) ectopic testes (Gray et al. 1993; Kelce et al. 1994); exogenous treatment of rats with the androgen dihydrotestosterone stimulates testicular descent (Frey et al. 1983). Treatment of male mice with DES on d 9-16 in utero led to an increase in the incidence of cryptorchidism (McLachlan et al. 1975). Male mammals lacking either adequate concentrations of androgen or androgen receptors exhibit a high incidence of cryptorchidism (Wilson and Foster 1985). Those data indicate that compromised (or altered) androgen receptors-because of receptor abnormalities or because of the presence of an antiandrogen or potent estrogen-preclude normal testicular descent during development in mammals. The elevated concentrations of p,p'-DDE present in the tissue of three female Florida panthers (Facemire et al. 1995) and the knowledge that this metabolite of DDT is known to exhibit antiandrogenic activity (Kelce et al. 1995) suggest a relationship between contaminant exposure and cryptorchidism in Florida panther cubs. However, the tissue of panthers has been shown to be contaminated with a variety of toxic substances, including mercury (Roelke et al. 1991). Electroejaculation studies of 12 male Florida panthers have shown low sperm density, poor sperm motility, and elevated numbers of sperm defects (Facemire et al. 1995). Sperm density (concentration/milliliter) averaged 4.8 ± 1.4 x 106 sperm in Florida panthers compared with 15.4 ± 4.4 x 106 for males from a Texas population and 22.5 ± 9.2 x 106 in Latin American populations (Barone et al. 1994). The male panthers studied had 24-50% more sperm abnormalities than were found in Texas panthers and significantly smaller testicular volumes (Barone et al. 1994). Furthermore, males had abnormal sex-steroid ratios, exhibiting higher concentrations of estradiol-17ß than testosterone in plasma (Facemire et al. 1995). As yet, there are no exposure data for other populations against which the contaminant concentrations in Florida panthers can be compared. Although the available evidence suggests that the reproductive anomalies could be the consequence of environmental contaminants, the role of extensive inbreeding in this small population cannot be discounted. Summary and Conclusions Several reproductive and developmental disorders have been observed in wildlife and human populations exposed to environmental contaminants, including HAAs. Laboratory studies using male and female rats, mice, and guinea pigs, and female rhesus monkeys have shown that exposure of these animals during development to certain HAAs (e.g., DDT, methoxychlor, PCBs, dioxin, bisphenol A, octylphenol, BBP, DBP, chlordecone, and vinclozolin) can produce structuralcontinue

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Page 168 and functional abnormalities of the reproductive tract. Some of these studies, according to the investigators, were conducted using doses at or near levels encountered in the environment, but in most instances the environmental relevance of the dose used is unknown, because of lack of data concerning the level of environmental contamination. With the exception of PCBs, TCDD, and DDT and its metabolite DDE, there are few human studies on the reproductive and developmental effects of exposure to HAAs. The effects of prenatal exposure to PCBs, DDE, and other contaminants from maternal consumption of contaminated fish or other food products has been studied in several populations in the United States and abroad. Collectively, these studies indicate that prenatal exposure to PCBs can cause lower birth weight and shorter gestation, and have also been correlated with IQ and memory deficits as well as delayed neuromuscular development. Pre- and post-natal exposure to PCBs and PCDFs from accidental contamination of rice oil in Yusho, Japan and Yu-Cheng, Taiwan have resulted in various developmental defects. Exposure of men to environmental HAAs has been suggested as the cause of worldwide increases in hypospadias, cryptorchidism, testicular cancer, and declines in sperm concentration. Studies examining these trends show considerable variation, both temporally and geographically. The degree to which the results reflect differences in the populations selected for study (fewer men of proven fertility, or men with concerns about their sperm concentration), diagnostic practice, or other methodologic differences, is the subject of continued controversy. With respect to the end point most closely studied, sperm concentration, retrospective analyses of trends over the past half-century remain controversial. When the data from large regions are combined together and analyzed, some data sets indicate a statistically significant trend consistent with declining sperm concentrations. However, aggregation of data over larger geographic regions may not be an appropriate spatial scale for this analysis given significant geographic heterogeneity in genetic and environmental factors. The current data are inadequate to assess the possibility of trends within more appropriately defined small regions. Acquiring data at smaller regional scales is critical to assessing the significant geographic variation in sperm concentration which is the subject of collaborative studies currently being conducted in the United States (funded by NIEHS), Europe (funded by the European Union), and Japan (funded by the Japanese EPA). Many wildlife studies show associations between reproductive and developmental defects and exposure to environmental contaminants, some of which are HAAs. One of the best established linkages between exposure to an environmental contaminant and reproductive effects in birds has been the correlation of DDT, and its metabolite DDE, with eggshell thinning. Many potential mechanisms for DDE-induced eggshell thinning have been described. The most current hypothesis is that the mechanism involves an inhibition of prostaglandin by the shell-soft

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Page 169 gland mucosa. However it does not appear likely that this is a result of DDE acting as a hormone receptor agonist or antagonist. Reproductive and developmental abnormalities have also been observed in several populations of fish exposed to effluents from sewage treatment plants and paper mills and polluted waters of the Great Lakes. Effects observed include intersexes in trout exposed to sewage treatment plant effluent (STPE); increased egg and fry mortality in Great Lakes trout and salmon; thyroid enlargement in Great Lakes salmon; and changes in plasma sex-steroid concentrations, decreased egg and gonad size, and delayed sexual maturity in whiter suckers exposed to effluents from paper mills along Lake Superior. Laboratory experiments with specific HAAs found in those effluents and polluted waters have produced effects consistent with these wildlife observations. For example, certain HAAs found in STPEs induce estrogenic responses in male trout. Specifically, ethinylestradiol and alkylphenol ethoxylates have been shown to induce vitellogenin synthesis, a hallmark of estrogen exposure, and to decrease the rate of testicular growth in male fish in tests that duplicate concentrations found in some effluents. Dioxin and structurally related compounds have been shown to induce blue sac disease in trout and reduced growth and survival of salmon. Thyroid enlargement in salmon of the Great Lakes is hypothesized to be caused by exposure to PCBs, which also have been shown to induce goiter formation in laboratory rodents fed PCB-contaminated salmon. Finally. B-sitosterol found in paper-mill effluent has been shown to alter the reproductive physiology of goldfish under experimental conditions. Laboratory studies are also consistent with some reproductive and developmental abnormalities (e.g., skewed sex ratios, behavioral modifications, and morphologic abnormalities of the gonads) observed among North American gull populations. Specifically, it has been shown that gull eggs injected with DDT at concentrations found in wild gull eggs induce gonadal abnormalities that are similar to those observed in contaminated gulls. Also, doves fed mixtures containing DDE and PCBs exhibit abnormal breeding behavior. Similarly, defects seen in alligators from Lake Apopka (the site of a chemical spill containing dicofol and DDT) including small penis size and abnormal testes in males and abnormal ovaries in females, are consistent with structural and functional reproductive abnormalities that occur following perinatal exposure of laboratory rodents to estrogenic and antiandrogenic chemicals. It has also been suggested that cryptorchidism. the most common reproductive anomaly found in male Florida panthers, is the result of exposure to p,p'-DDE. Because testicular descent is in part androgen dependent, and because antiandrogens and potent estrogens have induced cryptorchidism in rats and mice. it is plausible that exposure to contaminants with antiandrogenic or estrogenic properties could be causing the effects in male panthers. However, the Florida panther population is exposed to many other contaminants, including methoxy-break

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Page 170 chlor, PCBs, and mercury, and the role of extensive inbreeding in this small population cannot be discounted. Recommendations Based on evaluation of reproductive and developmental effects observed in humans, laboratory animals, and wildlife exposed to HAAs, the committee recommends that wildlife and human populations continue to be monitored for adverse developmental and reproductive effects. Specifically, the committee recommends the following: —Studies of wildlife that exhibit population declines, abnormal sociosexual behavior, or deformities should be designed to investigate those phenomena in light of specific environmental factors, including chemical contamination and environmental degradation. —Prospective and cross-sectional studies, using common protocols and strict quality control, be conducted in human populations suspected of being affected by HAAs. Serum hormone concentrations, body burdens of HAAs, and sperm concentration in seminal fluid should be measured, especially in relation to any adverse effects, or banked for later exposure assessment. Prospective and cross-sectional studies are particularly needed on cohorts tracked from conception through adulthood on female and male reproductive end points such as sperm concentration, cryptochidism, and hypospadias. —Regional differences in male reproductive end points such as sperm count and rates of hypospadias, cryptorchidism, and testicular cancer should be examined prospectively to determine whether the differences can be associated with genetic and environmental factors. Such prospective analyses should be accompanied by quantitative sensitivity analyses. —Free range farm animals should be studied for potential effects of environmental contaminants on fertility.break