nonylphenol and octylphenol) may be the causative agent. (Sumpter & Jobling 1995, Jobling et al. 1996). These products have been shown to be estrogenic in ...
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Commentary Endocrine disruptors and reproductive development: a weight-of-evidence overview R L Cooper and R J Kavlock
Introduction Recently, there has been considerable discussion in both the scientific literature and the lay press regarding the possibility that environmental chemicals, through their effects on endocrine function, are responsible for a number of reproductive and developmental anomalies in a wide range of wildlife species from invertebrates, through fish, reptiles, birds and mammals, including humans. The endocrine system consists of a number of central nervous system (CNS)-pituitary-target organ feedback pathways involved in the regulation of a multitude of bodily functions and the maintenance of homeostasis. As such, there are several target organ sites at which an environmental agent could disrupt endocrine function. Furthermore, because of the complexity of the cellular processes involved in hormonal communication, and the central role hormones play in regulating differentiation in early life stages, many of these functions could participate mechanistically in a toxicant’s effect, especially in developing organisms. In this Commentary we wish to discuss the weight of evidence that environmental chemicals are causing endocrine disruption in various species, particularly in response to exposures during critical developmental periods. For the purposes of this discussion, an endocrine disruptor is defined as an exogenous agent that interferes with the synthesis, storage/release, transport, metabolism, binding, action or elimination of natural blood-borne hormones responsible for the regulation of homeostasis and the regulation of developmental processes (Kavlock et al. 1996). Effects on development of the reproductive tract with clear linkage to an endocrine-mediated mechanism Currently, the most salient environmental culprits in the litany of ecological and health effects have been those compounds suspected of interfering with the normal action of estrogen through its receptor (i.e. estrogen agonists and antagonists). The fact that these hormones play a critical role in the normal development of the reproductive tract and sexual differentiation of the brain is well documented. Journal of Endocrinology (1997) 152, 159–166 0022–0795/97/0152–0159 $08.00/0
Thus, environmental compounds with estrogenic activity, such as kepone (Gellert 1978) and methoxychlor (Gray et al. 1985), and pharmaceuticals such as diethylstilbestrol (DES) (McLaughlin & Dixon 1977), have been shown to masculinize the female rodent’s brain, induce precocious puberty and disrupt cycling in adulthood. There is also clear evidence in humans that developmental exposure to estrogenic compounds can permanently alter reproductive tract development and physiology as evident in male and female offspring of mothers who were given DES during their pregnancy (Herbst et al. 1971, Gill et al. 1979). Disruption of sexual differentiation following exposure to estrogen has also been demonstrated in other species, such as the turtle, which show temperature-dependent sexual differentiation. Placement of either estrogen or some hydroxylated polychlorinated biphenyls (PCBs) that are estrogen agonists directly on the egg have been shown to alter sexual differentiation (Crews et al. 1995). Similar findings have been reported in birds (Fry & Toone 1981). Environmentally released chemicals may also have antiandrogenic properties. Compounds such as the vinclozolin metabolite M2 and the dichlorodiphenyltrichloroethane (DDT) metabolite, pp’-DDE, inhibit androgen binding to the androgen receptor (Kelce et al. 1994, 1995) and androgen-induced transcriptional activity (Wong et al. 1995). In vivo studies of vinclozolin and pp’-DDE have shown that these compounds inhibit androgen action in developing, pubertal and adult male rats (Gray et al. 1994, Kelce et al. 1995). Similar phenotypic observations have been reported in the alligator population of Florida’s Lake Apopka (Guillette et al. 1994). These alligators have a high incidence of altered sexual differentiation of the male reproductive tract and show feminized steroid hormone profiles, reportedly in response to a massive spill of the DDT-like pesticide dicofol into the lake in the early 1980s (Guillette et al. 1995a,b). Recently, concern has been expressed over the possibility that some man-made chemicals present in surface waters and aquatic sediments may adversely affect reproduction in fish (i.e. possible pseudohermaphrodite, smaller testes weight; Purdom et al. 1994, Sumpter 1995). In this case, the focus has been on the possibility that the
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biodegradation products of a major group of non-ionic surfactants, the alkylphenol polyethoxylates (especially nonylphenol and octylphenol) may be the causative agent (Sumpter & Jobling 1995, Jobling et al. 1996). These products have been shown to be estrogenic in vitro (Soto et al. 1991, White et al. 1994a), are found in high concentrations in the sediment around sewer effluents (Naylor et al. 1992, Ahel et al. 1994) and have been shown to bioaccumulate in several species (Ekelund et al. 1990). However, a true cause and effect relationship between these compounds and a population level impact in fish has yet to be demonstrated. Phytoestrogens can also influence reproductive function in fish. Gonadal weight and gonadal hormone levels in the serum of certain fish in the Great Lakes region residing downstream of pulp mills (Van Der Kraak et al. 1992), and in goldfish exposed similarly in the laboratory, are significantly reduced (MacLatchy & Van Der Kraak 1995). The fish downstream from the mills show a delay of several years in the time to sexual maturation. These investigators indicated that the levels of the wood-derived compound, â-sitosterol, found in high concentrations in the effluent of pulp mills, may be the primary agent responsible for these endocrine effects in fish (McMaster et al. 1991). Importantly, these observations demonstrate that altered endocrine or reproductive functions are not always the result of man-made environmental agents, but a consequence of an altered distribution of natural products in the environment. The observation that â-sitosterol has the potential to alter reproduction in fish is not unlike the problems of infertility observed in livestock (especially sheep) that graze on phytoestrogen-rich pasture or fodder, a situation that can be a major economic problem in many parts of the world (for review see Price & Fenwick 1985). Another important factor concerning the steroid receptor is that they are promiscuous with regard to the ligands with which they interact. That is, many of the chemicals classified as environmental estrogens can actually bind to more than one type of steroid receptor. For example, o,p-DDT and chlordecone can bind to the estrogen and progesterone receptors with similar affinities. Other compounds such as nonylphenol and the metabolite of methoxychlor, 2,2-bis-(p-hydroxyphenyl)-1,1,1trichloroethane, have the ability to inhibit binding to the estrogen, progesterone and androgen receptors (Laws et al. 1995). Effects on development of the reproductive tract with potential link to endocrine disruption While in the above examples there is a strong scientific basis for disruptions in the hormonal support of development or reproductive viability, other evidence of endocrine alterations following exposure to environmental compounds is more tentative, either from the standpoint of Journal of Endocrinology (1997) 152, 159–166
biological documentation of an endocrine-based mechanism, or from an evaluation of the criteria of causality for epidemiological studies. For example, exposure during gestation to 2,3,7,8-tetrachloro-p-dibenzodioxzin (TCDD) has been reported to cause malformations of the female reproductive tract (Gray & Ostby 1995), reduce ejaculated sperm counts (Mably et al. 1992a, Gray et al. 1995) and feminize the male’s brain (Mably et al. 1992b). These actions are presumably ultimately mediated via interaction of TCDD with the Ah (arylhydrocarbon) receptor, but careful examination of the effects indicates that they are neither purely estrogenic, anti-estrogenic, androgenic nor anti-androgenic in nature. While it is known that TCDD displays anti-estrogenic activity in some in vitro systems (Safe 1995) and in the adult female mouse liver (Goldstein et al. 1990), and has a variety of effects on growth factors in developing tissues (Abbott & Birmbaum 1990), the exact mechanism of action through which developmental disturbances are mediated in the reproductive tract remain obscure. In a similar vein, there are reported effects on wildlife species exposed to dioxins and other Ah receptor ligands. Because of the common receptor, the potency of mixtures of Ah agonists is represented collectively as toxic equivalency factors (TEFs), where the biological activity is expressed as a mass balance sum of the ratios of the potency of individual congeners (PCBs, polychlorinated dibenzofurans and polychlorinated dibenzodioxins) relative to the prototype 2,3,7,8-TCDD. Using the TEF approach, Peterson and co-workers (Cook et al. 1994, 1996) have completed a detailed description of the reproductive problems associated with lake trout in Lake Ontario. This species stopped natural reproduction in the lake in the 1940s. Laboratory studies using newly hatched trout showed that exposure to dioxin causes a marked increase in the incidence of blue sac disease and morphological anomalies, effects similar to those observed in natural populations. By using sediment analysis to reconstruct the contaminant levels in the water and survival of lake trout, they have shown that the survival rate is inversely proportional to the level of TEF contamination, which argues that such contamination is not compatible with the survival of the offspring. While the toxicity is unquestionable and the exposure reconstruction convincing as to the causative agent, the linkage of the mechanism to an endocrine alteration is absent. Like the effects of TCDD on mammalian development, it is problematic to describe them as endocrine disruption, even given the broad-scale definition put forth above. An equally tenuous association with endocrine disruption is present in reports correlating decreased offspring survival and PCB exposure in mink, in which it was shown that offspring survival rate decreases in a dose–response manner in animals fed fish diets containing increasing concentrations of PCBs (Heaton et al. 1995), and all other contaminants that may have been present in the fish.
Commentary
There are also a number of reports indicating that a disturbing decline in human sperm count, altered sperm morphology and sperm motility has occurred since the 1940s (Carlsen et al. 1992, Auger et al. 1995, Irvine et al. 1996), driving what is considered to be an already tenuous capability of the male to successfully produce viable offspring (Table 1). Accompanying this are concerns of an increase in the occurrence of developmental reproductive anomalies in humans, such as hypospadias and cryptorchism, and in testicular and prostate cancers in adulthood. Indeed, exposure of male laboratory animals to estrogenic (Gray et al. 1989, Sharpe et al. 1995) and anti-androgenic (Gray et al. 1994) compounds has been shown to result in lower sperm counts. It would seem equally logical to hypothesize that exposure of men during development to chemicals with similar mechanisms of action would lower their sperm count. Application of the principles of causality used in epidemiological research is a useful guidepost in the critique of the hypothesis about declining semen quality. Since no direct link between a particular environmental exposure and a decrement in semen quality is currently present, the strength of the association is necessarily weak, and no dose–response relationship can be demonstrated. In addition, the element of temporality of exposure is likewise absent and almost impossible to reconstruct given the presumed latency period. In terms of the specificity of the association, semen quality is affected by many factors, including smoking, temperature, season, diet, lighting, period of abstinence and concurrent diseases; thus it does not provide a pathognomic indicator to a particular insult. There is a moderate amount of consistency of effects across studies, although the Carlsen study and the Auger study appear to differ as to the trend in declining quality over the last several years, and at least three recent reports (Bujan et al. 1996, Fisch et al. 1996, Paulsen et al. 1996) do not support a downward trend in recent years. Given our knowledge of reproductive development, it is not beyond the realm of biological plausibility that environmental endocrine disruptors could induce the hypothesized effect in man (Sharpe & Skakkebaek 1993). However, we should also note that, although there are decrements in sperm counts in the DES cohort, the effect appears largely confined to cases with cryptorchid testes and there does not seem to be an effect on fertility in general in those men (Wilcox et al. 1995). The overall coherence of the evidence can therefore only be judged to be moderate at best (contrast with the above discussion of dioxin-TEFs and effects on lake trout reproduction in Lake Ontario). Therefore, the current state of the science makes any links with these disorders and environmental agents preliminary. Although the suggestions of a decline in male reproductive health raise serious concerns, data supporting these effects are equivocal. Exposure assessment is often lacking and the important associations between exposure and effect are made more difficult by the long latency
between exposure and observed effects. Furthermore, designing a prospective study to carefully evaluate the hypothesis is virtually beyond the realm of feasibility. Yet we are still faced with the possibility that the hypothesis is true and that steps may need to be taken to reduce additional impacts on the reproductive health of the human male. Hopefully, investigators will be able to identify existing cohorts of men that have documented and quantified exposures to suspected endocrine-disrupting chemicals during perinatal development and evaluate their reproductive health. Perhaps then we shall be able to reduce the uncertainty sufficiently either to take necessary actions or direct our attention to other aspects of the issue. Even more weakly supported associations between environmental contaminants and alterations in the reproductive system have been reported. For example, the recent account of altered sexual development in one beluga whale that died in the St Lawrence Seaway has been attributed to a high level of exposure to contaminants present in these waters. Although there has been some effort to document the level of exposure to environmental contaminants (especially PCBs) in cetaceans, in which significant concentrations of these chemicals have been reported to be present in their blubber, there are no clear examples of adverse effects in these mammals (White et al. 1994b). Perhaps the best example of confused cause-and-effect relationships lies in the case of egg shell thinning in raptors and other fish-eating birds by organochlorine pesticides (particularly DDT) first noted in the 1960s and 1970s. We now understand the primary mode of action is inhibition of the Ca-activated ATPases in the shell gland rather than the effect of these agents as either direct acting estrogens or as enhancers of metabolic turnover via induction of hepatic mono-oxygenase enzymes (Kolaja & Hinton 1977). Non-sex steroid-based effects on reproductive development or function It is clear from the above comments that the major focus in the area of endocrine disruptors has been the possibility that substances in the environment may interact with sex steroid receptors (primarily the estrogen receptor) to either mimic or inhibit the effect of the natural ligand. The function of virtually all the steroid receptor-mediated pathways can be disrupted by one or more environmental chemicals. For example, a number of compounds (particularly some PCBs; Goldey et al. 1995) have been shown to alter thyroid hormone activity. This is perhaps most evident in demonstrating that environmental agents can impair thyroid hormone-mediated events during development, leading to permanent alterations in brain function in adulthood. Not only have these observations been made in laboratory animals, but learning disabilities have been noted in children exposed to high levels of PCBs in either cord blood or breast milk (Rogan et al. 1987, Jacobsen et al. Journal of Endocrinology (1997) 152, 159–166
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Journal of Endocrinology (1997) 152, 159–166
1938–1991
14 947
34 (17–64)
Proven ‘fertile’ and normal men
40% decline between 1940s and 1990s
Meta-analysis; dependent variable was year of publication
Years covered
Sample size
Age of subject (range)
Source of subjects
Major findings for sperm count
Comments
na, not assessed.
Worldwide
Population source
Carlsen et al. (1992)
No change in volume; age increased during study (32–36); 3·3% decline in sperm count per year of age
2·1% decrease per year during study
Healthy, fertile, unpaid semen donors; 3–5 days abstinence
34 (15–59)
1351
1973–1992
Paris, France
Auger et al. (1995)
Volunteer donors to gamete research program; 3–4 days abstinence
27 (18–53)
577
1984–1995
Scotland
Irvine et al. (1996)
Positive association with year of birth; 3·3% increase in sperm concentration per year of age
Most recent donors averaged 14 years younger than donors born in 1950s
No change with time Later year of birth associated with reduced sperm count
Candidate fertile donors
34 (21–44)
382
1972–1992
Toulouse, France
Bujan et al. (1996)
Average sperm counts (#106) were: CA, 72·7; MN, 100·8; NY, 131·5); mean age increased during study (30–38); sperm count decreased 3% per year of age
Slight (0·65% per year) increase in sperm count over study
Prevasectomy donors; 3–10 days abstinence
34 (na)
1283
U.S. (NY, MN and CA) 1970–1994
Fisch et al. (1996)
na
5481
1967–1994
Turku, Finland
Vierula et al. (1996)
No change in sperm count over study
Mean age increased from 28 to 33 during study; older age contributed to decreased sperm count; mean count was 134#106/ml
No change in sperm count over study
Used geometric mean of multiple semen samples to reduce intra-subject variability
Healthy volunteers Clinical investigations in intervention study, of infertile couples; or all samples before 3–5 days abstinence clinical trials; 2–7 days abstinence; multiple samples
28 (18–52)
510
1972–1993
Seattle, WA
Paulsen et al. (1996)
Table 1 Summary of publications that have analyzed historical trends in human sperm counts. See text for weight-of-evidence discussion of the issue of causality
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1990). Interestingly, thyroid hormones are also involved in the regulation of testicular development. Increasing levels of thyroxine during the prepubertal period inhibit folliclestimulating hormone secretion and halt Sertoli cell proliferation. Almost paradoxically, testicular size and sperm production in rats is elevated by early neonatal inhibition of thyroxine production by propylthiouracil (Kirby et al. 1992) or Arochlor 1242 and 1254 (Cooke et al. 1996), as the prolonged period of hypothyroidism allows for continued Sertoli cell proliferation. Still other environmental agents have been shown to alter natural ligand synthesis or metabolism. These agents have the potential to interfere with endocrine function in many ways, and there are many examples in the literature documenting each case. Agents such as ketoconazole (Feldman 1986) and ethylene-1,2-dimethanesulphonate (Jackson & Morris 1977) have been shown to interfere with testosterone synthesis. â-Sitosterol, an estrogenic chemical based upon its ability to induce vitellogenin in male fish and an estrogen-like response in MCF-7 or T-47-D cells (Mellanen et al. 1996), has also been shown to lower gonadal steroid production in fish by possibly altering cholesterol availability or inhibiting cytochrome P450 activity. The dithiocarbamates (e.g. thiram, metam sodium and disulfiram) or compounds that may form dithiocarbamates in the cell (such as carbon disulfide) have been shown to inhibit norepinephrine (NE) synthesis in rodents, birds and man. The effect of brief exposures to these compounds on the hormonal control of ovulation, pregnancy outcome and behavior are well documented (Stoker et al. 1993, 1996). The chlorotriazine, atrazine, represents another class of compounds that disrupt the CNS control of pituitary-ovarian function. In rats, atrazine induces pseudopregnancy or prolonged diestrus depending upon the dose (Cooper et al. 1996a). This disruption of regular cycling is apparently the result of an inhibition of the pulsatile release of hypothalamic gonadotropinreleasing hormone (GnRH) and subsequently pituitary luteinizing hormone release (Cooper et al. 1996b). However, the exact hypothalamic mechanism(s) responsible for this effect on the releasing hormone has not been defined. This observation raises another interesting question concerning the mechanism(s) associated with chlorotriazine exposure and the development of mammary tumors in rats. It has been hypothesized that atrazine brings about an early onset of mammary gland tumors because it induces precocious aging within the neuroendocrine control of reproductive function. Specifically, it is argued that atrazine brings on an early onset of constant estrus and the endocrine milieu that is conducive to mammary tumor development (Stevens et al. 1994). However, the presence of repetitive pseudopregnancies and diestrus in atrazinetreated young adult animals is certainly not consistent with this hypothesis. The adverse effects of other compounds can be mediated through membrane receptors. Thus, exposure to
Figure 1 Environmental agents can alter endocrine function through a variety of mechanisms. Some of the better-studied cellular target sites include (1) steroid hormone receptor-mediated changes in protein synthesis and/or mitosis (e.g. methoxychlor, a weak estrogen, or DDE, an anti-androgen). Less well-studied cellular mechanisms include those alterations that occur after exposure to (2) compounds that interfere with membrane receptor binding (e.g. chlordimeform, an á-noradrenergic receptor blocker), (3) steroidogenesis (i.e. certain imidazole compounds) or (4) compounds that interfere with the synthesis of other hormones (e.g. dithiocarbamates disrupt adrenalin synthesis) and (5) compounds that alter the flux of ions across the membrane (e.g. pyrethroid insecticides alter sodium and chloride ion flux and metals compete for normal calcium ion flux) within certain types of hormone-secreting cells. Ca++=calcium ion, SH=steroid hormone in blood or cytoplasm, PH=peptide hormone, R=steroid hormone receptor, SRE=steroid response element, PK=protein kinase.
the formamidine acaricides, such as chlordimeform (which block the á2-NE receptor and disrupt the NE control of GnRH), induces a cascade of changes that include delayed ovulation, altered development in utero and fetal loss (Cooper et al. 1994). These are but a few examples and there is a large possibility that many other environmental agents may disrupt endocrine function through similar effects on CNS neurotransmitter and neuropeptide functions. In fact, there are a number of reports identifying the importance of catecholamine and cholinergic activation of normal sexual differentiation in rodents (Dohler 1991). Thus, just as exposure to estrogenic or androgenic compounds during certain periods is critical in reproductive development, exposure to compounds that may interfere with normal maturational changes in these neurotransmitter and neuropeptide systems may have similar permanent reproductive effects. Summary and conclusions It is clear that the endocrine system presents a number of target sites for the induction of adverse effects by environmental agents (Fig. 1). There are numerous Journal of Endocrinology (1997) 152, 159–166
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examples demonstrating that reproductive and developmental processes may be exquisitely sensitive to exposure and there are clear effects induced by presumed endocrine-disrupting chemicals in a variety of species. The concerns raised by studies of wildlife and humans place added significance on a better understanding of the myriad of effects attributed to endocrine disruptors. But there remains a large void between the study of relatively high exposure levels used in laboratory settings versus the relatively low levels found in the general environment. It is also equally clear that the term ‘endocrine disruption’ has been applied to situations where the biological basis is far from conclusive. This may be a moot point in situations where populations are experiencing adverse effects on reproduction, but as scientists we must be vigilant of the appropriate use of descriptive terminology, particularly in cases where public awareness and concern are as great as this. Not only is there a need for better test procedures (both in vivo and in vitro) to characterize the potential of environmental agents to disrupt endocrine function in laboratory species, but there is also a need for a more comprehensive understanding of the normal physiological processes associated with reproduction and development in those wildlife species studied. At the same time, obtaining better information on the transport, fate and bioavailability of chemicals released into the environment remains an important but imposing task. The goals of risk characterization are to carefully delineate cause-and-effect relationships, define the dose–response relationships, and determine whether environmental exposures exceed acceptable levels. A concerted research effort is needed to fill the voids in our knowledge and reduce the large uncertainties that exist today. Only then can regulatory actions take place within the confines of legislative mandates, remediation strategies and considerations of international use and transport. Towards this end, the US Environmental Protection Agency (EPA) sponsored two workshops in 1995 (Ankley et al. 1996, Kavlock et al. 1996) at which groups of international scientists began the process of identifying research needs. Similar efforts also took place in several European countries at about the same time (Danish Environmental Protection Agency 1995, Medical Research Council 1995, Umweltbundesamt 1995). More recently, a Working Group on Endocrine Disruptors has been established within the Committee on the Environment and Natural Resources of the US Government’s National Science and Technology Council. The objectives of this Working Group are to (1) formulate a framework for identifying research needs related to the health and ecological effects of endocrine-disrupting chemicals; (2) conduct an inventory of on-going federal research programs; and (3) identify research gaps and facilitate a co-ordinated research plan to address them. These efforts were largely completed in the Fall of 1996 and the information will be made available via the Internet Journal of Endocrinology (1997) 152, 159–166
(http://www.epa.gov/endocrine). The group also plans to work more broadly with other governments and private industry and public interest groups conducting research on this issue to co-ordinate research and disseminate scientific information. Persons wishing to know more about this effort should contact the authors. The issue of endocrine disruption has raised the consciousness of many researchers, both within and outside the toxicology community, and has attracted considerable public and political interest. We now have the beginnings of international co-operation to identify the most important scientific uncertainties and to dedicate resources to address the critical gaps. It is important that we develop rational approaches using the best science to determine the true risks that endocrine disruptors pose to humans and wildlife. The urgency of these efforts has been heightened by the recent passage of the Food Quality Protection Act of 1996 and the Safe Drinking Water Act of 1996 by the US Congress, both of which commit the US EPA to develop and implement a comprehensive screening program for estrogenic and other endocrine effects as deemed necessary by the Administrator. EPA must develop this screening effort within 2 years of enactment, implement it within 3 years, and report to Congress in 4 years, clearly an ambitious schedule. Disclaimer The research described in this article has been reviewed by the National Health and Environmental Effects Laboratory, US EPA, and approved for publication. Approval does not signify that the contents necessarily reflect the views and policies of the Agency, nor does mention of trade names or commercial products constitute endorsement of recommendation for use. References Abbott BD & Birnbaum LS 1990 TCDD-induced altered expression of growth factors may have a role in producing cleft palate and enhancing the incidence of clefts after coadministration of retinoic acid and TCDD. Toxicology and Applied Pharmacology 106 418–432. Ahel M, Giger W & Koch M 1994 Behavior of alkylphenol polyethoxylate surfactants in the aquatic environment – I. Occurrence and transformation in sewage treatment. Water Research 28 1131–1142. Ankley GT, Johnson RD, Toth G, Folmar FC, Defenbeck NE & Bradbury SP 1996 Development of a research strategy for assessing the ecological risk of endocrine disruptors. Reviews in Toxicology B Environmental Toxicology (In Press). Auger J, Kunstman JM, Czyglik F & Jouannet P 1995 Decline in semen quality among fertile men in Paris during the past 20 years. New England Journal of Medicine 332 281–285. Bujan L, Mansat A, Pontonnier F & Mieusset R 1996 Time series analysis of sperm concentration in fertile men in Toulouse, France between 1977 and 1992. British Medical Journal 312 471–472. Carlsen E, Giwereman A, Keiding N & Skakkebaek NE 1992 Evidence for decreasing sperm quality of semen during the past 50 years. British Medical Journal 304 609–613.
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Reproductive Toxicology Division, National Health and Environmental Effects Research Laboratory, US Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA
