Human Reproduction Update Advance Access originally published online on December 10, 2007
Human Reproduction Update 2008 14(1):59-72; doi:10.1093/humupd/dmm025
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Impact of endocrine disruptor chemicals in gynaecology
1 Institute of Gynecology, Perinatology and Child Health, Sant'Andrea Hospital, University of Rome La Sapienza, Via di Grottarossa 1035, 00189 Rome, Italy 2 Department of Food Safety and Veterinary Public Health, Istituto Superiore di Sanità, Rome, Italy
3 Correspondence address. Tel: +39 6 33775696; Fax: +39 6 3350611; E-mail: donatella.caserta{at}libero.it
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Abstract |
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The potential hazardous effects that estrogen- and androgen-like chemicals may have both on wildlife and human health have attracted much attention from the scientific community. Endocrine disruptors (EDCs) are chemicals that have the capacity to interfere with normal signalling systems. EDCs may mimic, block or modulate the synthesis, release, transport, metabolism and binding or elimination of natural hormones. Even though potential EDCs may be present in the environment at only very low levels, they may still cause harmful effects, especially when several different compounds act on one target. EDCs include persistent pollutants, agrochemicals and widespread industrial compounds. Not all EDCs are man-made compounds; many plants produce substances (phytoestrogens) that can have different endocrine effects either adverse or beneficial in certain circumstances. Natural substances such as sex hormones from urban or farm wastes can become concentrated in industrial, agricultural and urban areas; thus, such wastes may be considered potential ‘EDCs’ for humans and/or wildlife. Much attention has focussed on changing trends in male reproductive parameters in relation to EDC exposure; however, studies on the female reproductive system have been less comprehensive. We have focussed this article on four major aspects of female reproductive health: fertility and fecundability, endometriosis, precocious puberty and breast and endometrial cancer.
Key words: environmental effects / endometriosis / female infertility / estrogen
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Introduction |
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TOPAbstractIntroductionConclusions and recommendationsReferences |
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Both in humans and in animals, endocrine signalling is involved in reproduction and embryo development, growth and maturation, energy production, use and energy storage, electrolyte balance and maintenance and behaviour. Hormones trigger such complex functions by interacting with their receptors that are present, at a nuclear and/or cellular level, in various organs and tissues as part of a complex biological feedback system. Any disruption of this balance can cause impairment in the physiological status of the whole organism, especially during the more susceptible developmental stages. If the regulatory role of the endocrine system is impaired, abnormal function and/or development of the reproductive, the nervous and the immune systems may occur. It has been recently postulated that predisposition for certain types of tumours is caused by an altered development during prenatal growth (intrauterine) and in the first years of life (Sharpe and Sakkebaek, 1993
). It is noteworthy that the available data overall support the public health concerns for endocrine disrupting chemicals (EDCs). Several substances, including EDC, with share the ability of interfering with the female reproductive system and possibly implicated in the development of some gynaecologic pathologies are listed in Table 1.
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This review article summarizes the available literature on adverse outcomes in female human reproductive health from supposed exposure to EDCs, or from which an endocrine mechanism is plausible or has been proven.
An endocrine disruptor is an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub)population (WHO, 2002). The homeostasis of sex steroids and the thyroid are the main targets of EDC effects; hence, reproductive health, considered as a continuum from gamete production and fertilization through to intrauterine and post-natal development of progeny, is recognized as being especially vulnerable to endocrine disruption (Mantovani, 2002). EDCs are widespread in food chains and in the environment and include persistent organic pollutants (POPs) such as the insecticide dichloro-diphenyl-trichloroethane (DDT) and its metabolites, the industrial by-product dioxins and the industrial compounds polychlorinated biphenyls (PCB), several agrochemicals, pesticides and biocides (e.g. chlorinated insecticides, organotins, imidazoles and triazoles) and other industrial compounds (several phenol compounds such as bisphenol A) (Mantovani et al., 1999). More recently, attention has been focussed on the endocrine effects of substances (e.g. parabens, component of UV-screen, phthalates) widely diffused also in cosmetics and toiletries (Oishi, 2002; Kunz and Fent, 2006), as well as of metals, e.g. arsenic compounds (Tseng et al., 2003). Substances other than typical food and/or environmental contaminants may also be considered as EDCs, such as drugs, anabolic agents and especially phytoestrogens: this class of chemicals includes isoflavones, lignans, etc., which are present in some food items such as soy, and in cosmetics with active ingredients of vegetal origin. A good dietary intake of phytoestrogens might act as protective factor against several cancers (e.g. breast, prostate) and post-menopausal diseases (e.g. osteoporosis); however, there is some concern regarding the exposure to high doses during pregnancy or early infancy, e.g. through the use of dietary integrators or soy-based milk (Skibola and Smith, 2000; Stark et al., 2003), due to the high intake of hormone-mimic substances during the critical period of infant and adolescent development. Many EDCs are known to act as agonists of estrogen receptors (ER), e.g. bisphenol A and alkylphenols, or to antagonize androgen receptor (AR) such the dicarboximide fungicides; progesterone receptors (PR) are also a potential target for many chlorinated EDC, such as DDT and derivatives (Scippo et al., 2004). Compounds that act as hormone triggers should also be taken into account; dioxins and the dioxin-like chemicals being the best-known examples. They bind to the cytoplasmic aryl hydrocarbon receptor (AhR), which in turn cross-talks with the steroid nuclear receptors to initiate entirely new responses; these may vary with the status (activated or not) of the ER and AR. Other chemicals such as some phthalate plasticizers or the polybrominated diphenyl ethers (PBDEs) used as flame retardants are more likely to interact with less specific orphan nuclear receptors, like pregnane-X or androstane, which act as ‘sensors’ to regulate the activity of ligand-specific receptors (Sanders et al., 2005; Wyde et al., 2005). Atrazine and related herbicides instead can impair the hypothalamus-hypophysis-gonads axis (McMullin et al., 2004). Other EDCs can interfere with hormone synthesis and transport, examples are the azole antifungals agents, used in agriculture and animal production, that inhibit different steps of steroid synthesis (Zarn et al., 2003) and the hydroxylated PCB metabolites that selectively interact with estrogen sulphotransferase, thus increasing estradiol (E2) bioavailability in target tissues (Kester et al., 2000). Such distinctions however should not be considered too rigidly; recent evidence indicates that the type of effects induced by several EDCs may vary with sex and age of the target organism (Pryor et al., 2000; Hallgren and Darnerud, 2002). The adverse effects of xenobiotics on female reproductive health, namely fertility and pregnancy maintenance, have already been identified in a number of studies (Hruska et al., 2000). Epidemiological evidence points to known lifestyle factors (cigarette smoke, high alcohol consumption and obesity) and intensive occupational exposures to heat, radiation or solvents as risk factors for reproductive dysfunction in women, although with variable incidence (Eggert et al., 2004; Hassan and Killick, 2004; Kumar, 2004).
The initial interest for the effects of EDC on human health was related to the possible interference with male reproductive disorders (Sharpe and Sakkebaek, 1993). However, there is growing evidence from toxicological as well as clinical studies that the female reproductive system is a potential EDC target too.
Hazard identification—mechanisms of EDC actions
Experimental data on animal models indicate that early prenatal and/or perinatal exposure to EDCs can lead to long-term effects on reproduction and development which can become evident later, even at sexual maturity and/or at adulthood. The identification and characterization of this ‘early exposure—late effect’ pattern of EDCs still represents a challenge for scientists and risk assessors. Accordingly, there is an ongoing international effort to develop, validate and/or update the current toxicological testing approaches. To screen substance for their potential effects on endocrine homeostasis, the 28-day oral toxicity study [Organisation for Economic Co-operation and Development (OECD) Guideline 407], i.e. the common subacute toxicity test widely applied to all chemical testing strategies, has been enhanced to detect potential EDC through additional parameters, including: serum hormone concentrations, estrus cycle features and weights of endocrine-related organs (ovaries, uterus with content, etc.) (Kunimatsu et al., 2004). Although it is a comprehensive test to identify compounds with endocrine activities, the assay retains possible limitations for the full characterization of effects and dose-response relationships, including the relatively short duration and the testing of adult animals.
The two-generation reproductive toxicity study (OECD Guideline, 416) represents the most effective test to evaluate alterations on the endocrine homeostasis during the entire developmental and reproductive period. Since human exposure to EDCs may span a lifetime, multigenerational studies are usually chosen so that the test substance is administered continuously, without interruption, to parental (P) and subsequent offspring generations (F1, F2, etc.). This protocol will provide information about effects on male and female reproductive performances, potency and fertility, pregnancy outcomes, maternal lactation and offspring care, prenatal and post-natal survival, growth and development of offspring, as well as their reproductive capacity.
The above-mentioned experimental protocol requires a large number of animals and is also costly, laborious and time consuming. Nevertheless, at the moment it represents the only reliable tool for the study of such a complex system and/or function, e.g. reproduction and to perform hazard identification for EDCs.
There are a number of mechanisms whereby EDCs can modulate endocrine systems and potentially cause adverse effects on human health (Fig. 1). The generally accepted paradigm for receptor-mediated responses includes hormone binding to its receptor at the cell surface, cytoplasm or nucleus, followed by a complex series of events that lead to changes in gene expression (Birnbaum, 1994). The main nuclear receptors involved in EDC action are: ER 
and β, AR, thyroid receptors, AhR, glucocorticoid receptors (GR), as target for arsenic (Bodwell et al., 2004) and pregnane-X-R, which appears to be specific target for some phthalates (Hurst and Waxman, 2004). More recently, attention has been focussed on PR that appear to be more sensitive than ER- and a target for many POPs (Villa et al., 2004). Thus, a panel of in vitro assays for binding/transactivation of nuclear receptors may be highly relevant for screening and identification of potential EDCs. Unfortunately, this is likely to be insufficient, as several EDCs do not specifically interact with nuclear receptors.
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Other relevant mechanisms include inhibition of hormone synthesis, transport, or metabolism and activation of receptor through receptor phosphorylation or the release of cellular complexes necessary for hormone action. In the case of hormone synthesis, considerable research has been conducted on the aromatase inhibitors; in fact they can prevent the conversion of androgens to estrogens through a cytochrome P450 system, which is highly conserved among the species. Several azole fungicides have been shown to cause aromatase inhibition as well as some widespread organotins (tri-butyl tin and tri-phenyl tin) (Matsui et al., 2005
). In addition, there is a growing awareness that multiple receptor systems act in concert to regulate biological functions. For example, ‘cross-talk’ between the ERs and growth factor receptor appears necessary for estrogen signalling in mammary cells to divide or differentiate. These events are critical for explaining several risk factors for breast cancer, such as age at menarche, age at menopause or effects on the number of pregnancies.
Other example of a multi-factored EDC-mediated effect is the hypothalamo-hypophysis-gonadal axis (HHG). Atrazine, the well-known broadleaf herbicide, despite its limited solubility, is commonly found in the ground and in surface water together with its biotransformation by-products, creating concern for exposure to the general population. Atrazine exerts pleiotropic reproductive effects in rodents, which comprise alteration of estrous cycles, delay of the onset of puberty and mammary gland development in prenatally and/or post-natally exposed females, increased pre and/or post-implantation losses, induction of mammary tumours as well as acceleration of reproductive ageing (Friedmann, 2002; Rayner et al., 2004). All these effects are thought to be related to a more direct alteration of HHG axis, in particular to the LH suppression caused by inhibition of GnRH and prolactin (Rayner et al., 2005).
Thus, a sensitive, time- and cost–effective evaluation of the thousands of chemicals in use that have not yet been tested for their endocrine disrupting potential requires the development of an integrated battery of assays encompassing different mechanisms and targets (Bremer et al., 2005).
Dose–response relationships
Dose–response relationships is perhaps the most controversial issue regarding EDCs. One of the reasons is that EDCs often act by mimicking or antagonizing the actions of naturally occurring hormones. These hormones are already present at physiologically functional concentrations, so the dose–response considerations for EDCs are often different than for other chemicals acting throughout different pathways. In particular, some estrogenic compounds can induce an inverted-U dose–response curve, resulting from low-dose stimulation of response; this kind of behaviour challenges current methods of risk assessment. Several estrogenic compounds, such as bisphenol-A, octylphenol can induce effects at doses lower than those inducing general toxicity; these adverse stimulatory inputs divert energy needed for other processes resulting in reduced and/or altered health performances. The current threshold models used for the assessment of low doses of chemicals showing this shape of dose–response curve tend to underestimate the risk and deserve more attention (Weltje et al., 2005).
A common dose–response for all endocrine disruption mechanisms should not be expected. This conclusion is based on the knowledge that chemicals categorized as endocrine disruptors have shown many different mechanisms of action. These activities include estrogenic, antiestrogenic, antiandrogenic, growth factor modulation, cytokine and thyroid modulation, modulation of hormone metabolism. Besides, epidemiological documentation of endocrine disruption is complicated when exposures are mixed. In fact, even if the endocrine activity (estrogenic and androgenic, etc.) and the mechanisms of all compounds were known, the risk of multiple exposures can not be assessed from a single-chemical approach as used in traditional toxicology. Although mixtures of compounds can have effects, it may not be possible to attribute causation to a single chemical. Furthermore, for EDCs, cumulative low-dose insult can be more toxic than a single high-dose exposure; this tends to modify the hypothesis of classical toxicology. Two possible approaches have been proposed to study mixtures: the investigation of mixtures without specifying the individual components; and the study of the individual compounds with a focus on possible interactions in the context of dose to the receptor (Koppe et al., 2006). Thus, analysis of mixtures allows evaluation of combined effects of chemicals each present at low concentrations (Rasmussen et al., 2003).
Exposure assessment
Several chemicals in the environment (e.g. pesticides, industrial chemicals and natural products) have shown to be hormonally active; they can be detected and measured in wildlife, in environmental samples as well as in the human population. POPs are persistent in the environment; they are lipophilic and they can bioaccumulate in fat-rich tissues, such as milk, dairy products and fatty fish (Jacobs et al., 2002). Breastfeeding provides a significant source of exposure to POPs early in human life, the effects of which are not completely understood yet; nevertheless, despite of the possibility of harm from environmental contaminants in breast milk, breastfeeding is still recommended as the best infant feeding method (Nickerson, 2006). Other compounds may be less persistent in food and environment but can be more toxic, above all if the exposure falls in a critical period of the life cycle, e.g. development or maturation. Knowledge about the magnitude of human or wildlife exposure is often limited. Nevertheless, the presence in the food chain of environmental chemicals other than POPs may be more important than it was assumed a decade ago. Examples include: the presence of nonylphenols, detergent by-products with estrogenic properties, in seafood (Ferrara et al., 2001); the POP-like behaviour of widespread industrial chemicals such as PBDE (Schecter et al., 2005) and perfluorooctane compounds (Olsen et al., 2005). A major issue is the assessment of real exposure, i.e. internal exposure levels through biomonitoring. Information is available only for some compound groups (e.g. pesticides) in workplaces (Aprea et al., 2002). Recently, increasing attention is paid towards bio-monitoring of the general population not only for metals, POPs and POP-like compounds but also for different EDCs. For example, pilot studies conducted in Italy recorded high levels of phthalates in the umbilical cord blood of neonates and in the serum of women with endometriosis (Cobellis et al., 2003; Latini et al., 2004), highlighting that the general population is currently exposed to low-doses of EDCs. Similarly, US studies on urinary metabolites of phthalates, confirm the widespread presence of these substances in the environment (Calafat and McKee, 2006).
The definition of appropriate biomarkers of exposure for monitoring is also critical. In fact, for some EDCs such as phthalate acid esters, alkylphenols, ethylene bis-dithiocarbammate fungicides, PCBs, PBDEs, phytoestrogens, it is important to evaluate the relative exposure and toxicity of the main metabolites which represent the real active compounds at the target site (Elsby et al., 2001).
Exposure assessment at a community level is a complex issue that involves usage patterns of the different compounds and how they are transported and metabolized in different environmental compartments (water, sediment and biota). Exposure often needs to be considered specifically for vulnerable groups, such as communities with special dietary habits (e.g. fishing communities) (Yeats et al., 1999); lifestyles may modulate intake of contaminants, e.g. smokers have consistently higher levels of cadmium, a heavy metal which is included among potential EDC (Castelli et al., 2005). Children deserve special consideration as they have relatively greater food and water intakes, breathing volume and contact with soil as compared to adults, leading to a potentially higher uptake of chemicals per kilogram of body weight (Schwenk et al., 2003).
Risk characterization—factors related to differential vulnerability to EDCs
It is well known that differential responsiveness to EDCs has been observed among different species as well as inside a general population. The biologic and molecular mechanisms underlying this specificity are quite diverse.
Determinants of species-specificity include differences in receptor binding, gene transcription patterns of gene expression and cellular responses to endocrine-active compounds (Gray et al., 2004). Inside general populations, differences in responsiveness may be determined at the level of genetic polymorphisms in hormone-metabolizing enzymes, hormone receptors and those genes that are activated by these receptors (Li et al., 2005). Our rapidly growing knowledge—emerging from the human genome project—will enable the design of rational studies on the impact of EDC exposures on hormonally sensitive endpoints in groups that may be genetically predisposed. Thus, it is important to identify genetic factors influencing the individual and/or population's susceptibility to endocrine disorders and to set appropriate biomarkers (Masse et al., 2002). Differences in the response to EDCs rely also on the life cycle phase of exposure. Children, e.g. show a specific vulnerability and sensitivity linked to the immaturity of organ/tissues as well as metabolic systems (LaRonda et al., 2004). Extrinsic factors such as diet and lifestyles can also impact individual susceptibility to endocrine-active agents. Vulnerability of different groups in the population will be affected by lifestyle factors, in particular diet: several experimental studies indicate that the effects of toxicants may be modulated by the intake of, e.g. phytoestrogens (You et al., 2002) and antioxidants (Kocdor et al., 2005; Muthuvel et al., 2006).
Therefore, more attention to the nutrient-toxicant interactions should be given in epidemiological studies. The use of new approaches in molecular toxicology and epidemiology as well as more targeted experimental protocols have the potential to yield additional valuable information to elucidate the role of these mechanistic determinants of specificity at low-dose exposures to potential EDCs and to improve risk evaluation for the adverse health effects of EDCs.
EDCs and women's reproductive health
Fertility and fecundity
Fertility and fecundity have shown a progressive decrease in the last decades (Hassan and Killich, 2004). Occupational exposure is often cited as a risk factor for female fertility, as well as for early pregnancy loss and pre-term delivery. Pesticides represent a relevant example: direct exposure through pesticide handling may be included among specific risk factors for reproductive health in the rural environment. In one case–control study, in which 281 women with a diagnosis of infertility were compared with 216 post-partum women, those women with a history of working in the agricultural industry showed an elevated risk of infertility (Fuortes et al., 1997). de Cock et al. (1994) found that reduced fecundability ratio and longer time-to-pregnancy are associated with application of pesticides, in particular when older spraying techniques were used. Greenlee et al. (2003) identified an association between adverse reproductive outcomes in women and the practice of mixing and applying herbicides as well as the use of fungicides. Greenhouse work is a peculiar type of agricultural work that implies a continuous exposure to pesticides. In a large Danish study on time-to-pregnancy among female workers in flower greenhouses, the overall fecundability rate did not differ between workers and referents. However, certain factors, possibly associated with a specifically increased exposure to chemicals, are consistently associated with a significant 20–30% reduction of fecundability, i.e. handling cultures many hours per week, spraying of pesticides and the lack of glove use (Abell et al., 2000). A major problem in this sort of study is the very general consideration of pesticide classes (e.g. ‘fungicides’ and ‘herbicides’), whereas these include highly diverse chemical groups, with only a few being EDC or reproductive toxicants: a detailed appraisal of compounds specifically related to adverse reproductive outcome is indeed important for prevention and risk communication strategies. An attempt to identify exposures to specific pesticides has been represented by the Ontario Farm Family Health Study. In a retrospective study on 2012 farm couples, no strong or consistent pattern of association between exposure to various classes of pesticides and time-to-pregnancy was observed (Curtis et al., 1999). Data from this large study also suggested that female reproductive dysfunctions (e.g. reduced fecundability and increased risk of miscarriage) may be associated to exposure to pesticides of the male partner. In fact, miscarriage risk increased with reported use of several compounds, including thiocarbamates, atrazine and phenoxy herbicides, especially when more compounds were used at once and protective equipment was not used (Savitz et al., 1997; Arbuckle et al., 1999). The possible role of male partner exposure was supported by an Italian retrospective studies that showed significantly increased risks of conception delay and spontaneous abortion among the spouses of high-exposure workers (greenhouse, applicators), even after adjusting for possible confounders such as age, education, smoking habits, etc., the risk increased with reported exposure to some pesticide groups only, including potential EDC such as chlorinated insecticides and triazines (Petrelli et al., 2000, 2003).
Regarding the general population exposed through food or the living environment, a few studies relate the possible EDC exposure to markers of impaired reproduction. The US Great Lakes are considered a problem area concerning long-term pollution to persistent EDCs, mostly PCB.
In the New York State Angler Cohort Study, lifetime exposure to PCBs, estimated taking into account the fish consumption from contaminated Great Lakes, was correlated to changes in fecundability (Buck et al., 1999). Maternal consumption of fish for 3–6 years was associated with reduced fecundability [odds ratios (OR), 0.75; 95% confidence interval (CI), 0.59–0.91). This effect was of the same magnitude in those with >7 years of fish consumption, however, it did not reach statistical significance (OR, 0.75; 95% CI, 0.51–1.07). Studying a Michigan cohort (Courval et al., 1999), increasing potential lifetime exposure to PCBs and mercury, estimated by the numbers of sport fish meals consumed, was associated with an increased time-to-pregnancy related to increased paternal consumption; the association reached statistical significance only in the group with highest fish consumption (adjusted ORs were 1.4, 1.8 and 2.8 for 1–114, 115–270 and 271–1127 annual fish meals consumed, respectively).
Studies that have associated adverse reproductive effects together with biomarkers of exposure are of great interest. A positive association between the body burden of persistent chlorinated EDC (PCB, DDT and metabolites, other chlorinated insecticides) and gynaecological problems, especially recurrent miscarriages, was observed in case–control studies carried out in Germany (Gerhard et al., 1998) and USA (Korrick et al., 2001). In a limited pilot study, maternal exposure to the estrogen-like bisphenol A has also been associated with recurrent miscarriages (Sugiura-Ogasawara et al., 2005); more studies are required to support or refute this interesting finding. Nevertheless, conflicting observations do exist in regards to the role of EDC exposure: a study on Japanese patients with a history of recurrent miscarriage has not found an association with serum levels of industrial products such as PCBs, hexachlorobenzene (HCB) or DDE (Sugiura-Ogasawara et al., 2003). Thus, possible mechanisms related to different vulnerability deserve more attention. In this respect the possible influence of dietary phytoestrogens should also be considered. Several studies have shown prolongation of the menstrual cycle in healthy premenopausal women with a daily intake of 45 mg of isoflavones from soy protein as possibly attributable to a prolongation of the follicular phase for the suppression of the normal midcycle surge in FSH and LH (Cassidy et al., 1994; Lu et al., 2000). Nevertheless, in a study on flaxseed ingestion, the luteal phase was reported to be prolonged (Phipps et al., 1993). Since flaxseed is especially rich of lignans, the possibility that different phytoestrogens do exert diverse actions cannot be ruled out.
Overall, apart from evidence from ecological studies, data on the actual exposure to the EDCs impairing female fertility are limited. The isolation of persistent organochlorine chemicals from ovarian follicular fluid of women undergoing IVF at the moment could represent a useful tool to obtain information using the available reproductive technology techniques (Jarrell et al., 1993a,b). Isolation of such chemicals at a critical period of oocyte development may provide an important biomarker of exposure.
Endometriosis
Endometriosis is an estrogen-dependent disease characterized by the presence of endometrial glands and stroma outside the uterine cavity. It is a common gynaecological disorder as well as a major cause of infertility (Chedid et al., 1995) affecting ~14% of women of all reproductive ages (Vercillini et al., 1995). At present, it is generally recognized that there is no one single theory to identify and explain all aspects of that multi-factored clinical syndrome. Although there is a clear association with estrogens, it is accepted that the disease is not specifically caused by estrogens but is stimulated by them (Guarnaccia and Olive, 1998).
Previous work in non-human primates has shown that exposure to the dioxin 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD, the ‘Seveso’ dioxin) is associated with an increased prevalence and severity of endometriosis. Rodent studies support the possible role of environmental contaminants in the pathophysiology of endometriosis, although a convincing mechanistic hypothesis has yet to be advanced (Rier et al., 2003). In cynomologous monkeys treated with TCDD at dose levels of 0, 1, 5 and 25 ng/kg body weight/day for 12 months, the growth of surgically implanted endometrial fragments was increased at the two higher dose levels circulating gonadal steroid levels and menstrual cycle characteristics were unchanged (Yang et al., 2000).
Serum levels of TCDD and dioxin-like PCB were analysed in 13 years monkeys after termination of a 4-year study with dietary exposure to 0, 5 and 25 ng/kg diet. Elevated serum levels of total dioxin-like compounds correlated strongly with endometriosis. (Rier and Foster, 2001). These findings may be relevant to humans since the serum levels recorded in TCDD-exposed animals were similar to concentrations reported in serum, milk and tissues from the general population. Interestingly, TCDD-exposed rhesus monkeys with endometriosis exhibited long-term alterations in systemic immunity, such as enhanced tumour necrosis factor-alpha secretion by blood monocytes (Rier et al., 2001).
A relationship between endometriosis and exposure to dioxin-like PCBs (Gerhard and Runnebaum, 1992) and dioxins (Koninckx, 1999) has been proposed. A positive association between endometriosis and dioxin exposure was reported in a case–control study in which 44 women with endometriosis were compared with 35 age-matched controls with tubal infertility (Mayani et al., 1997). Significantly more women with endometriosis [8 (18%) versus 1 woman (3%) of the control group] had detectable dioxin levels in their serum (P = 0.04); there was no relationship between severity of endometriosis and concentration of dioxin. In another case–control study, no association between plasma organochlorine concentrations and endometriosis could be found in 86 women with endometriosis compared with 70 controls, matched for the indication of laparoscopy (Lebel et al., 1998). The highly exposed women in Seveso also underwent an evaluation for endometriosis in a case–control study, which showed no significant association between dioxin levels and the presence or amount of endometriosis, although a trend towards increased risk was apparent in the group with the highest body burden (>100 ng/l serum); the authors cautioned that disease misclassification in a population-based study may have led to an underestimate of the true risk of endometriosis (Eskenazi et al., 2002).
The plasma concentrations of ubiquitous environmental contaminants such as phthlates are associated with endometriosis in an Italian study that for the first time suggests the role of phthalate esters in the pathogenesis of the disease (Cobellis et al., 2003).
There are no direct data regarding the impact of phytoestrogens, such as isoflavones or flavanones, on ectopic human endometrium, although it could be anticipated that they would cause effects, given that soy isoflavones inhibit E2-mediated endometrial proliferation in macaque monkeys. Hence, the human data at present have failed to provide compelling evidence for or against an association of EDCs and endometriosis.
Precocious puberty
Environmental chemicals have also been suggested as potential causative factors in the temporal modification in the age of puberty onset. Population-based studies showing a reduction in the median age of puberty demonstrate the need for investigations of the possible causes of this trend (Delemarre-van de Waal, 2005).
A hypothesized external agent that induces central precocious puberty would act through the early initiation of the pulsatility of gonadotropin-releasing hormone from the hypothalamus, thereby inducing the cascade of hormonal events that result in pubertal development. However, review of the literature has not yet established a conclusive relationship between central precocious puberty and environmental agents. In the case of peripheral precocious puberty, the mechanism in young women is mediated by hormonal receptors in peripheral tissues that are responsive to estrogen or estrogen-like compounds.
In Puerto Rico, a temporal trend toward premature breast development (premature thelarche) in girls and gynaecomastia in boys has been noted during the early 1980s (Mills et al., 1981; Bongiovanni, 1983; Saenz de Rodriguez et al., 1985; Freni-Titulaer et al., 1986). Serum samples from 41 girls from Puerto Rico with premature breast development and 35 control cases were analysed for determining the possible presence of pesticides and/or phthalate esters. No pesticides or their metabolites could be found in any of the serum samples; however, 28 (68%) of the girls with premature breast development had measurable levels of phthalates [dimethyl, diethly, dibutyl and di-(2-ethylhexyl)], compared with 6 of the 35 (17%) control samples (Colon et al., 2000). In another study, performed in Belgium, plasma of girls with precocious puberty was screened for chlorinated pesticides; data showed an increase in plasma levels of the DDT metabolite p,p'-DDE in foreign children with precocious puberty immigrating from developing countries to Belgium, compared with undetectable levels in Belgian-born girls with idiopathic or organic precocious puberty; since DDT is still used in a number of developing countries, the data suggest a possible relationship with early exposure to this pesticide (Krstevska-Konstantinove et al., 2001).
The effect of in utero exposure to polybrominated biphenyls (PBBs) on sexual maturation was evaluated in Michigan girls whose mothers were accidentally exposed through diet to the flame retardant FireMaster (Blanck et al., 2000). Effects on pubertal end points were assessed by questionnaires sent to mothers of daughters <18 years of age and to the daughters themselves. The data reveal that menarche and pubertal hair growth were significantly advanced in girls breastfed, i.e. with higher perinatal exposure to PBBs. Namely, the adjusted OR (CI) were 0.8 (0.3–1.9) for early menarche in the non-breastfed and 3.4 (1.2–9.0) for the breastfed group, respectively; for early pubertal hair growth they were 0.9 (0.2–4.3) in the not breastfed and 19.5 (2.8–138.2) in the breastfed, respectively. There was no association with Tanner stage of breast development. These findings are interesting because menarche and breast development are estrogen-dependent events whereas pubertal hair growth is independent from estrogen levels, suggesting that PBBs may interact with different pathways. Several experimental data showed alterations in morphological parameters of reproductive development following exposure to environmental xenobiotics both in animal and human studies (Laws et al., 2000a,b). For example, the chlorinated insecticide Metoxichlor, an EDC with estrogenic activity, is able to accelerate the vaginal patency in prepubertal female rats (Masutomi et al., 2003).
Due to the scarcity of human data with appropriate biomarkers, toxicological studies could be useful for a reliable risk evaluation; in particular, protocols focussed on prepubertal evaluation of EDCs could provide relevant data on a phase of specific susceptibility for human maturation and development (Kim et al., 2002).
Breast and endometrial cancer
The incidence of breast cancer increased steadily from the 1940s through the 1990s in many industrialized countries, with the highest risk found in Western Europe and North America. The increases may be explained, only in part, to the increased screening procedures.
Dietary factors are known to contribute to breast cancer risk (Willett, 2001). One class of dietary compounds that has received much attention is phytoestrogens. Consumption of phytoestrogens, particularly soy products, is higher in Asia than in the Western World (Messina et al., 1994), where the rate of breast cancer is highest (Kelsey and Berstein, 1996). In Japan and China, breast cancer rates are half as high as in the USA (Coleman et al., 1993). Migrants from Asia to the USA typically acquired a breast cancer risk associated with their host nation by the second generation, suggesting a direct influence of the environment rather than a genetic factor (Barnes, 1998; Lamartiniere, 2000).
Extensive human studies support an etiological role for estrogens, all related to increased lifetime exposure to endogenous estrogen (Hulka and Stark, 1995). Serum concentration of 17β-E2 is ~40% lower in Asian women than their Caucasian counterparts (Peeters et al., 2003). In two studies on premenopausal women (Lu et al., 2000; Kumar et al., 2002), modest consumption of soy products seems be associated with a reduction of steroid hormone levels, whereas a year-long dietary intervention trial involving 34 premenopausal women fails to confirm this data (Maskarinec et al., 2002). The reduction in steroid hormone levels by phytooestrogens is proposed to occur via the direct regulation of 17β-E2 biosynthesis and metabolism (Limer et al., 2004).
Despite their apparent effect on endogenous hormone levels, the role of phytooestrogens in breast cancer initiation and development is unclear. In vitro studies in estrogen-dependent human mammary epithelial cells (MCF-7) have shown that genistein, the main soy isoflavone, at low concentrations (0.1–10 mM) can stimulate cell proliferation whereas at higher concentrations (
10 mM) can act as inhibitor (Miodini et al., 1999). Other studies suggest that both proliferative and antiproliferative effects might be observed, depending on tumour cell type, concentrations, timing of phytoestrogen exposure and type of phytoestrogen given (Aldercreutz and Mazur, 1997). This may be explained by the multiple mechanisms of action that phytoestrogens seem have in humans. The chemical structure of the isoflavones is similar to that of estrogens, these substances diffuse through the cell cytoplasm and bind to both the nuclear ER subtypes ER- and ER-β, although they preferentially bind to and activate ER-β; for this reason, they are sometimes classified as selective ER modulators (SERMs) (Brzezinski and Debi, 1999; Diel et al., 2004).
In vitro, phytoestrogens also possess a variety of non-hormonal properties. Several phytoestrogens, except genistein, in addition to their estrogenic effect also suppress the activity of the aromatase enzymes, which are responsible for conversion of androgens to estrogens (Almstrup et al., 2002). At concentrations in excess of 25 µmol/l, phytoestrogens are capable of inducing apoptosis in human breast cancer cells (Po et al., 2002), also in ER-negative cell lines, confirming the hormone-independent mechanism of action. Dietary phytoestrogens are capable of inhibiting the tyrosine kinase activity, that are involved in a number of growth factor signalling pathways, implicated in control of cell growth and differentiation. Furthermore, phytoestrogens have antioxidant activity (Ruiz-Larrea et al., 1997) and may stimulate the immune system (Zhang et al., 1999) and inhibit angiogenesis (Su et al., 2005).
In addition to the potential protective effects against breast cancer, some data suggest that phytoestrogens could actually promote breast cancer. In vitro and in animal studies, genistein stimulates the growth of estrogen-sensitive mammary cancer cells (Hsieh et al., 1998; Allred et al., 2001).
Epidemiological studies on the relationship between soy consumption and risk of breast cancer are also discordant. A Japanese prospective study conducted in a cohort of 35 000 women (Key et al., 1999) revealed no significant association between soy consumption during adulthood and breast cancer incidence. Similar data was found in a retrospective analysis in a multiethnic cohort of non-Asian breast cancer patient and control individuals residing in USA (Horn-Ross et al., 2001). Epidemiologic evidence suggests that breast cancer chemoprevention by dietary phytoestrogen compound might be dependent on ingestion before puberty, when the mammary gland is relatively immature (Shu et al., 2001). A meta-analysis of currently available studies indicates that high soy intake might reduce the risk of developing premenopausal breast cancer but has no effect on post-menopausal breast cancer risk (Trock et al., 2000). This raises the question of whether foods rich in phytoestrogens may have complex actions in such diseases as breast cancer, exerting both preventive and promoting effects and perhaps depending on whether or not the tumour is estrogen dependent. Published literature regarding the effects of ingestion of dietary phytoestrogens by breast cancer patients and survivors is, also, controversial (Messina et al., 2001; This et al., 2001).
Thus, caution is necessary in promoting the putative beneficial effects of phytoestrogens with respect to the overall mammary neoplasms (Bouker and Hilakivi-Clarke, 2000). The majority of data relating environmental EDCs to human breast cancer are limited to persistent organochlorine compounds that have been identified worldwide in human tissue, blood and milk (Adami et al., 1995).
Since the metabolites of DDT have been identified in serum, adipose tissue and breast milk of individuals with no history of occupational exposure and/or living in areas where DDT has not been used for years, long-term exposure to DDT through food could be hypothesized to increase the risk for developing estrogen-dependent tumours such as breast cancer. However, prospective case–control studies have failed to demonstrate that DDT, and more specifically p,p-DDE, increases the risk for breast cancer (Krieger et al., 1994; Hunter et al., 1997; Hoyer et al., 1998; Dorgan et al., 1999; Helzlsouer et al., 1999; Ward et al., 2000; Wolff et al., 2000a,b). In addition, two retrospective case–control studies performed with post-menopausal women (among which breast cancer tends to be more estrogen dependent) did not find increases in breast cancer risk related to DDT exposure (vant'Veer et al., 1997; Moysich et al., 1998), nor did other retrospective case–control studies including pre- and post-menopausal women (López-Carrillo et al., 1997; Dello Iacovo et al., 1999; Mendonca et al., 1999; Zheng et al., 1999; Aronson et al., 2000; Bagga et al., 2000; Demers et al., 2000; Millikan et al., 2000; Stellman et al., 2000; Wolff et al., 2000,b), in contrast to two positive studies (Olaya-Contreras et al., 1998; Romieu et al., 2000). A recent combined analysis of five US studies showed no relationship between p,p'-DDE and breast cancer risk (Laden et al., 2001a,b).
Many of the studies on the relationship between PCB exposure and breast cancer have involved occupational exposure and have not found a positive association (Adami et al., 1995; Houghton and Ritter, 1995; NRC, 1999). In a prospective case–control study in USA, after 9.5 years of follow-up, higher serum levels of HCB induced a 2-fold increased risk of breast cancer. However, there was no evidence for a dose-response relationship, and the association was limited to women whose blood was collected close to the time of diagnosis (Dorgan et al., 1999). In a recent Belgian study, the serum levels of DDE and HCB were compared in 231 women at the time of breast cancer discovery and in 290 age-matched healthy controls. p,p'-DDE was found in 76.2% of cases and in 71.1% of controls but HCB was present only in 12.6% of cases (29 from 231) and in 8.9% of controls (26 from 290). Mean serum levels were significantly higher in patients versus controls for both compounds (2- and 3-fold higher for DDE and HCB, respectively). No excess was observed among nulliparous women or when familial history of breast cancer was considered. In the cancer group, no differences in serum levels of p,p'-DDE or HCB were found in relation with ER status, Bloom stage or lymph node metastasis, but the HCB level was significantly correlated with tumour size (P = 0.026) (Charlier et al., 2004).
Some studies deserve attention as they were performed in highly polluted areas. The Seveso Women's Health Study comprises 981 women, who were infants to 40-years old in 1976 and resided in the most contaminated areas and had archived sera that was collected soon after the infamous chemical accident in 1976. Modelling serum TCDD levels in 15 women that were diagnosed with breast cancer, showed that individual body burden is significantly related with breast cancer risk (Warner et al., 2002). The pleiotropic effects of such potent hormone trigger as TCDD were in fact associated with a number of long-term health effects in Seveso exposed population, such as increased morbidity and mortality from lymphoemopoietic and other neoplasms as well as markers of altered endocrine-immune function; interestingly the increases were often gender related. One evident, albeit still difficult to interpret, effect was the clear link between exposure levels in men and a lowered male/female sex ratio in their offspring (Pesatori et al., 2003).
In the severely polluted Eastern Slovakia areas, a retrospective study on serum samples of 24 breast cancer patients and 88 population controls did show trends for positive correlation of breast cancer with DDE and for negative correlations with PCB and HCB; however, results were generally not significant, due to low statistical power (Pavuk et al., 2003).
Although epidemiologic studies have not shown a strong relationship between blood levels of PCBs and the risk of breast cancer, an US cohort study revealed a stronger association among post-menopausal white women with the inducible M2 polymorphism in the cytochrome P450 1A1 gene (Laden et al., 2002). In a further population-based case-control study, breast cancer risk is evaluated in relation to PCBs and the CYP1A1 polymorphisms M1 (also known as CYP1A1*2A), M2 (CYP1A1*2C), M3 (CYP1A1*3) and M4 (CYP1A1*4). The study population consisted of 612 patients (242 African American, 370 white) and 599 controls (242 African American, 357 white). The results confirm that CYP1A1 M2-containing genotypes modify the association between PCB exposure and risk of breast cancer and additional evidence suggesting that CYP1A1 M3-containing genotypes modify the effects of PCB exposure among African American women (Li et al., 2005a,b).
Moreover, recent data suggest that underarm cosmetics might be a cause of the development and progression of breast cancer, because these cosmetics contain a variety of substances with endocrine activity (e.g. parabens) that are applied frequently in an area directly adjacent to the breast (Darbre, 2006).
Uterine cancer is more common in developed countries, with a similar pattern of hormonal risk factors as breast cancer. There is clear evidence that unopposed estrogen is the major risk factor for endometrial cancer. Neonatal treatment of mice on post-natal days 1–5 with either the potent estrogen agonist diethylstilboestrol (DES) or the phytoestrogen, genistein, has been shown to cause uterine adenocarcinoma by 18 months (Potischman et al., 1996; IARC, 1999). As with other cancers, the timing of exposure is critical to the potential development of uterine cancer; in fact treatment of adult mice with comparable levels of DES does not induce uterine neoplasms (Newbold et al., 1991). Soy isoflavones inhibit E2-mediated endometrial proliferation in macaque monkeys (Cline and Foth, 1998). This may not be surprising in the case of phytoestrogens that have both estrogenic and antiestrogenic effects, acting as SERMs (Whitten and Patisaul, 2001).
Epidemiologic data on the effects of environmental EDCs on endometrial cancer are limited. Sturgeon et al. (1998) found no association between endometrial cancer and 27 PCB congeners, 4 DDT-related compounds and 13 other organochlorine compounds. Several retrospective occupational cohort studies also observed no association (Bertazzi et al., 1987; Brown, 1987; Sinks et al., 1996). In the Seveso industrial accident, TCDD exposure appeared to reduce the risk of uterine cancer, but the number of cases was too small for a comprehensive evaluation (Bertazzi et al., 1993).
There is some evidence that dietary isoflavones protect from endometrial proliferation. Specifically, high consumption of soy products and other legumes in US women was associated with a decreased risk of endometrial cancer for the highest compared with the lowest quartile of soy intake (Goodman et al., 1997; Horn-Ross et al., 2003). Controversially, a recent randomized doubled-bind, placebo-controlled study on 298 post-menopusal women shows an increased incidence of endometrial hyperplasia following 5 years of treatment with 50 mg of soy isoflavones (Unfer et al., 2004). Thus, phytoestrogenic supplements should be reconsidered, particularly in women at high risk for endometrial cancer.
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The major limiting factor in drawing any conclusions about female reproductive system effects and EDCs is the scarceness of actual exposure data. In fact, in the available studies, exposure can only be inferred and not actually measured. Another problem relies on the sample sizes which are often too small to allow detection of an effect, even if one were present. Although there is evidence for geographical differences and temporal trends in some aspects of human reproduction, there has been no systematic attempt to look for evidence that the mechanisms behind these changes could involve endocrine pathways. Despite these drawbacks, the biological plausibility of possible damage to human reproduction derived from exposure to EDCs seems strong when viewed against (i) the background of known influences of endogenous and exogenous hormones on many of the processes involved, and (ii) the evidence of adverse reproductive outcomes in laboratory animals exposed to EDCs. Thus, concerns about females derive more from biological knowledge about the influence of sex hormones on development and adult reproductive function, rather than from studies on environmental chemicals. Contrary to male fertility, data on EDC effects on female reproductive health are sparse both in the human and experimental literature; there is still no attempt of an unified hypothesis such as the ‘testicular dysgenesis syndrome’ elaborated for EDC-related effects on male reproductive system (Sharpe, 2003
). Evaluation of possible links to EDCs could be pursued by exploiting identified temporal trends and geographical differences in the sex ratio, whereas the frequency of endometriosis in the general population offers opportunities for human studies requiring fewer numbers than might be required for other end points. Time-to-pregnancy can also be used as an apical tool to broadly examine the possible influence of environmental chemicals on both male and female reproductive function.
Endocrine-related cancer and especially breast cancer are most interesting issues. Although numerous human epidemiological studies have been conducted to determine whether environmental EDCs may contribute to an increased risk of breast cancer, the results remain inconclusive. Overall, the current scientific evidence (from human and animal studies) does not support a direct association between exposure to environmental EDCs and increased risk of breast cancer, although the evidence may be stronger in situations of high pollution (Seveso, Slovakia) and in a fraction of genetically susceptible individuals (Ursin et al., 1997; Brunet et al., 1998; Jernstrom et al., 1999; Nkondjock et al., 2006). Such studies prompt the need for translational research on EDC-related mechanisms, leading to the characterization of biomarkers of exposure, effects and susceptibility in order to perform a reliable risk evaluation. Breast cancer is likely due to many factors, including genetics, lifestyle, diet, endogenous hormone status and environmental factors. Research on whether potential complex interactions among these factors, modulated by individual genetic susceptibility, produce breast cancer is critical. Until consistent and compelling data on these issues become available, the role of EDCs in breast or uterine cancer incidence is likely to remain a highly controversial issue.
However, all the studies published to date have measured EDC exposure levels in adult women. The time of development when exposure takes place may be critical to define the dose–response relationships of EDCs for breast cancer as well as for other health effects such as endometriosis. The perinatal period and the period between age at menarche and age at first full-term pregnancy may be particularly important for breast tumour development and latency. Data are also required to investigate the late health outcomes of precocious peripheral puberty, for which EDC may be also a risk factor.
The claim that the time of life when exposure takes place (e.g. prenatal, neonatal, childhood and adolescence) may be the most critical factor is supported by human data and by basic research in animal models. Regarding endocrine-active compounds, the case of DES is the most well-known example, with young adult offspring exposed in utero to this potent drug having a higher rate of reproductive tract abnormalities in both sexes as well as of the rare clear-cell vaginal adenocarcinoma in female offspring (Swan, 2000). Of course, most EDC are unlikely to be nearly as potent as DES; even in such a case, the exposure levels are expected to be much lower than those occurring from pharmacological treatment. Nevertheless, it cannot be excluded that adult women currently at risk for breast cancer may have been exposed to exogenous EDCs in utero or during infancy, childhood and adolescence in the mid-1900s when contaminant levels of organochlorines were higher. Experimental evidence shows that gestational and/or lactational exposure to dioxin-like chemicals or to bisphenol A alters the differentiation of mammary gland in rodents (Fenton et al., 2002; Muto et al., 2002; Munoz-de-Toro et al., 2005). Research is urgently needed to address the role of timing of exposure. Human prospective studies would be complex, time-consuming and expensive; researchers should be encouraged to utilize and develop animal models to address the important issue of late health effects of early exposures, e.g. prepubertal animal studies for the effects of EDCs on sexual maturation. In parallel, biological banks should be exploited in order to conduct retrospective follow-up and nested mother-child cohort studies using specific biomarkers.
In conclusion, the currently available human data are inadequate to support a conclusion about whether the female reproductive system is adversely affected by exposure to EDCs; however, the weight of the evidence is adequate to address further studies as well as to prompt precautionary actions against excess exposure to xenobiotics specifically active on hormonal homeostasis.
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