1Nutrition and Development, The Rowett Research Institute, Greenburn Road, Bucksburn, Aberdeen, AB21 9SB, UK; ²Obstetrics and Gynaecology & ³Medical Genetics, University of Aberdeen, Foresterhill, Aberdeen, AB25 2ZD, UK

Running title: EDCs and male reproductive health

Keywords: development/EDCs/review/testis/human

Correspondence and proofs to: Dr Paul A. Fowler, Dept. Obstetrics and Gynaecology, University of Aberdeen, Foresterhill, Aberdeen AB25 2ZD, Scotland, UK, Tel: (+44) 1224 652633, Fax: (+44) 1224 684880, Email: p.a.fowler@abdn.ac.uk

Acknowledgements: The authors thank Prof. Jorma Toppari for critical review of the manuscript and the James Alexander Mearns Trust for financial support.

1. Introduction

One in six couples in Britain are infertile and this has traditionally been regarded as a problem that women suffer from. However in many cases the male partner is the principle contributor to the couple's infertility (1,2). Reports of rapidly declining sperm counts (3-11), rising incidences of testicular cancer (12) and increases in male reproductive tract abnormalities (13,14) have also contributed to the groundswell of interest in male reproductive health over the last decade.

These concerns are centred around reports showing that our environment is contaminated with thousands of man-made chemicals that can interact with the endocrine system of humans and animals. These so-called endocrine disrupting chemicals (EDCs) are thought to mimic or block the action of hormones and therefore disrupt sexual development in utero. It is hypothesised that this disruption may explain the increased incidence of male reproductive dysfunction (15,16). Furthermore, in wildlife populations urinogenital tract malformations and reproductive diseases analogous to those observed in the human have been associated with the unintentional exposure to harmful chemicals. More recently animal models have shown that maternal EDC exposure induces reproductive dysfunction in male offspring (17-29). Even in the human similar abnormalities occur in the sons of mothers exposed to the synthetic estrogen diethylstilbestrol (DES) during pregnancy. All of these studies have raised considerable concern regarding the impact of environmental EDCs on the developing male fetus and subsequent adult reproductive health.

This review assesses the evidence for an association between environmental EDCs and an increased incidence of male reproductive dysfunction. Human epidemiological and in vitro data are discussed in combination with several animal models of in vivo EDC exposure. Possible pathways by which EDCs may affect male reproductive development are considered.

2. Wildlife observations
Many wildlife species have suffered a decline in male reproductive health and this decline has been extensively reviewed elsewhere (15,16). In many cases, this has been associated with environmental pollution and some of the key observations are summarised in Table 1 (30-33). For example, it has been suggested that the developmental peculiarities of gonads and abnormal sex hormone concentrations observed in juvenile male alligators from Lake Apopka, Florida, were caused by contamination of the lake following an extensive spill of pesticide (Table 1) (30). Reptiles have temperature dependant sex determination (34) which can be modified by estrogen treatment. For example turtle eggs incubated at 26^oC produce male offspring. However, if the eggs are "painted" with estrogenic chemicals and incubated at the same temperatures, females are produced (35). Alligators (36) and turtles (37) are both susceptible to this phenomenon and the data from Lake Apopka indicated that the reproductive function of these species had been irreversibly altered causing reduced circulating steroid levels and impaired sexual differentiation.

More recently it was reported that exposing male fish to the effluent from sewage treatment works induced the synthesis of vitellogenin in the liver (Table 1) (31). This protein, the precursor of yolk, is usually produced in females in response to estrogens from the ovaries. In contrast, males do not normally produce vitellogenin unless exposed to supra-physiological levels of estrogens (38). Indeed, vitellogenin synthesis has been associated with the feminisation of male fish in British rivers contaminated with sewage effluent. Although the mechanism is uncertain, deleterious effects on testicular structure and the cytology of both germ and Sertoli cells have been reported (39). Furthermore, very high concentrations of vitellogenin are associated with kidney failure and increased mortality rates (40).

Researchers studying herring gull colonies along the Pacific coast of the United States in the 1970's noted changes in their breeding behaviour, notably nests contained twice the normal number of eggs. This was found to be due to female birds sharing nests together instead of with males (Table 1) (32). Furthermore, the feminisation of male birds, which leads to a skewed male/female ratio (Table 1), has been associated with exposure to DDT (dichlorodiphenyltrichloroethane) and it’s main metabolite DDE. It was unclear whether feminisation of males or high mortality rates in this sex produced the skewed sex ratio (41) although the impaired fertility rates due to female-female paired nests were of concern (32).

Panthers native to Florida are an endangered species. Although this is predominantly due to destruction of their natural habitat, it has also been shown that their reproductive function has been affected in terms of a decrease in numbers of viable offspring (Table 1) (33). Adult male panthers show low semen volume and number and this is compounded by poor semen motility and a high percentage of morphologically abnormal sperm. The male cubs exhibit a high incidence of uni- and bi-lateral cryptorchidism indicative of perturbed development in utero (Table 1). The body fat in female panthers of breeding age has high levels of various contaminants including DDT and it’s metabolites which are thought to be derived from their major food source, the racoon (42). Therefore, exposure of panthers to this chemical legacy in utero is believed to be causing developmental abnormalities in cubs leading to impaired reproductive health in adulthood.

Recent studies of wild otter populations in the Columbia and Fraser River systems of northwestern North America have revealed that they exhibit abnormal sexual development (43). Researchers found that delayed or inadequate development of the reproductive tract of male otters was negatively correlated with levels of polychlorinated bisphenyls (PCBs) in the liver. Heavily contaminated males had shorter baculum lengths and smaller testes than less contaminated individuals (Table 1) (43). Furthermore, researchers studying European otter populations reported that average PCB levels in otters appear to be highest in areas where the species was in decline and thriving otter populations were correlated with low mean PCB tissue concentrations (44,45). However, high levels of contaminants have recently been found in thriving otter populations in Scotland, especially Shetland (44). Smit and colleagues point out that different otter populations may have different contamination profiles of PCB congeners i.e. the total PCB concentration does not reflect differences in individual congener toxicity (see section 3.3).

All of the above effects reported in wildlife populations have been attributed to exposure of the developing fetus to environmental pollutants. Since the reproductive abnormalities observed parallel those abnormalities reported to be increasing in the human, it has been hypothesised that the decline in male human reproductive health is due to exposure to a maternal store of environmental pollutants during human fetal life.

3. Chemicals with endocrine disrupting effects
The man-made chemicals with known endocrine disrupting effects can be classified into five groups: (a) organochlorine pesticides, (b) dioxin compounds, (c) polychlorinated bisphenyls (PCBs), (d) alkylpolyethoxylates (APEs) and (e) plastic additives. In addition, phytoestrogens occurring naturally in foods have known endocrine disrupting effects. It is currently estimated that there are at least 70 chemicals in the environment which are known to be reproductive endocrine disruptors (46). However, of the tens of thousands of chemicals in the environment, only a handful have been properly tested and those with known endocrine disrupting properties have largely been discovered by chance.

3.1 Organochlorine pesticides
Organochlorine pesticides with known estrogenic activity include dichlorodiphenylethanes (DDT and it's metabolite p,p'-DDE), dieldrin, methoxyclor and toxaphene. Many of these were used extensively until the 1960's when their use was banned or restricted in developed countries. However, because of the continued widespread use of DDT in the developing world, especially in malarial control, it is estimated that current worldwide use of DDT exceeds levels used historically in the West (47).

3.2 Dioxin compounds
The polychlorinated dibenzo-p-dioxin (PCDD) family consists of 75 congeners of which the most toxic is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (48). Dioxins, formed as by-products in the manufacture of chlorinated hydrocarbons, are also inadvertently produced during incineration and emitted by motor vehicles. Dioxins are very environmentally persistent and lipophilic, hence levels are highest in animals at the top of the food chain (49).

3.3 Polychlorinated bisphenyls (PCBs)
The PCB family includes 209 congeners used as additives in many products including hydraulic fluids, fire retardants and lubricating oils, until a ban on their use in 1977. These congeners are widespread and persistent in the environment, and are present in trace amounts in surface soils, air, vegetation and water. Studies have shown PCBs to be abundant in certain foods and human tissues (50,51). Similar to organochlorine pesticides, these compounds are lipophilic and bioaccumulate in the body fat (52). Both estrogenic and anti-estrogenic effects have been reported for different PCB congeners (53) and this is thought to be dependent on the degree of chlorination (54). Interestingly, chlorination has been shown to be an important factor in the successful transfer of PCBs across the placenta: less-chlorinated congeners were transported more readlily than the highly chlorinated PCBs (55). This property could explain why high total PCB levels have been associated with both stable and declining otter populations (see section 2).

3.4 Alkylphenol polyethoxylates (APEs)
Alkylphenol polyethoxylates (APEs) are widely used as industrial surfactants in detergents, paints, herbicides, pesticides and cosmetics. After microbial breakdown during sewage treatment, approximately 60 percent of APEs are released into the aquatic environment as short chain APEs such as octylphenol and nonylphenol (56-58). APEs are relatively persistent and lipophilic resulting in bioaccumulation in the lipid of living organisms (59,60).

3.5 Phthalate compounds
Phthalate compounds are used in the plastics industry to give products flexibility. Phthalates are the most abundant man-made environmental pollutants, consequently human intake by various routes is tens of milligrams per day (61). High levels have been detected in plastic wrapped foods due to phthalate compounds leaching from plastics into the high fat processed foodstuff. Although phthalate compounds only act as weak estrogens on MCF-7 breast cancer cells compared to octylphenol or DDT (62), continuous exposure to moderate levels in the environment could result in significant effects.

3.6 Bisphenol A
The lacquer coating on food cans, bisphenol-A, is a plastic monomer that has been shown to have an estrogenic effect on MCF-7 breast cancer cells (63). The same compound is used in plastic dental sealant and in both of these applications bisphenol-A has been shown to leach from it's original starting place. Bisphenol-A has been detected in tinned foods (64) and in the saliva of dental patients even six months after treatment (65).

3.7 Phytoestrogens
Many plants and fungi contain naturally occurring phytoestrogens (66) and humans are probably exposed to milligram quantities every day. Active substances are isoflavones (genistein and daidzen), coumenestans (coumesterol) and the fungal metabolite zearalenone. Legumes are also rich in phytoestrogens, with the highest concentrations found in soybeans and soybean based products. These foods contain approximately 0.2 - 1.6 mg of isoflavones per gram dry weight (67). Soy is being used with increasing frequency in the processed food industry as a relatively cheap source of protein. Therefore changes in diet in many countries over the past 50 years may have increased daily intakes of phytoestrogens.

In contrast to manmade EDCs, phytoestrogens are readily metabolised and excreted (68) and therefore do not bio-accumulate in the body fat. Indeed phytoestrogens may in some cases be beneficial (Table 2) and have been shown to ameliorate certain conditions associated with EDC exposure. There is evidence that consumption of a high phytoestrogen-containing diet may give some protection against breast, colon and prostate cancers (69). Furthermore, phytoestrogens may reduce the potency of EDCs by stimulating production of sex hormone binding globulin (SHBG) by the liver and hence decreasing the bio-availability of endogenous estrogen (70). Paradoxically, reproductive disorders in sheep grazing on red clover which is rich in genistein, have been widely documented (71), therefore ingesting large quantities of phytoestrogens could be deleterious in some circumstances.

4. Evidence for disrupted human reproductive health

4.1 Semen quality
Several reports have suggested a possible decline in human semen quality over the last 50 years (3-11). However, many of these studies have been discounted as having little relevance to the general population as they were based on data from men attending infertility clinics or from highly selected groups of fertile men (3,5,8,10,11). In 1992 an extensive meta-analysis of 61 studies that included 14947 apparently fertile men revealed that the mean sperm density had decreased from 1135 10⁶/ml in 1940 to 665 10⁶/ml in 1990 (7). This meta-analysis incorporated consideration of changing semen analysis methods over the time period. In addition, a significant reduction in seminal volume was also recorded. This definitive publication has stimulated an extensive debate with counterclaims of no change or even an increase in semen quality. Despite the inherent problems associated with comparing data over five decades, the general conclusion was that an overall real decline in sperm quality has occurred (16). This conclusion prompted researchers to analyse data from other countries regarding trends in semen quality over recent years. Hence, in a French study of 1351 volunteers from a Parisian sperm bank the mean concentration of sperm decreased by 2.1 percent per year, from 895 10⁶/ml in 1973 to 605 10⁶/ml in 1992 (5). In the same study the percentages of morphologically normal and motile spermatozoa decreased by 0.6 percent and 0.5 percent respectively whereas there was no change in semen volume. After accounting for age and duration of sexual abstinence, it was concluded that the semen quality of a fertile man in 1992 was significantly poorer than the equivalent individual in 1973. Analysis of data from Greece also revealed a significant decline in sperm number from 154.35 10⁶/ml in 1977 to 130.15 10⁶/ml in 1993 (3) accompanied by a decrease in semen volume although this did not achieve significance. Similar conclusions have been drawn from smaller studies carried out in Scotland (8), Belgium (10) and Denmark (11). More recently, a re-analysis of Carlsen’s meta-analysis (7) controlling for abstinence time, age, percent proven fertility, specimen collection method, study goal and location reported the sperm count decline to be greater in the USA and other Western countries than previously estimated (72). This publication prompted a National Institutes for Health press release linking the sharp decline in sperm density with increased exposure to EDCs and increasing incidences of testicular cancer (73). Furthermore, an unselected population of young Danish men was found to have a high frequency of sub-optimal semen quality with 40 percent of the men having sperm counts below 405 10⁶/ml (4). In contrast, semen analyses conducted in some parts of the United States and Finland show no decline in either quality (74,75) or number (76) of sperm. However, it is important to remember that certain wildlife populations exhibit geographical differences in reproductive health which cannot be correlated with EDC exposure and remain unexplained (44).

4.2 Testicular cancer
There is no doubt that the incidence of testicular germ cell cancer is increasing world-wide (12). More specifically, in Denmark the incidence of testicular germ cell cancer has increased 3-4 fold since the 1940's (77). Analysis of cancer registries in countries around the Baltic Sea revealed that birth cohort was an important determinant of testicular cancer risk (78) and therefore it was suggested that major etiological factors need to operate early in life, perhaps even in utero. Interestingly, many geographical locations reporting an increasing incidence of testicular cancer also report a decline in sperm quality. In contrast, Finland exhibits both a stable sperm count and the lowest incidence of testicular cancer (12). However, even in Finland a similar upward trend in testicular cancer ratio has been observed (79). In total, the geographical linkage between testicular cancer and reduced sperm count suggests a possible shared environmental cause.

All germ cell cancers, except for spermatocytic seminomas, appear to arise from cells of carcinoma in situ (CIS) within the testis (80,81). CIS cells differentiate from primordial germ cells possibly during fetal life although the stimulus for this is unclear. An increased risk of developing testicular cancer has been reported in several groups of men including gonadal dysgenic (82,83), infertile (84) and low birth weight (85) individuals. In brief, all conditions that are reported to increase risk of testicular cancer are associated with a relative delay of intrauterine development or an imbalance in the activity of sex hormones during fetal development. It has been suggested that an abnormal cellular environment may enhance the survival of early germ cells which then persist in their undifferentiated form and subsequently transform into CIS cells. Together with the dramatic increase in prevalence of testicular cancer in specific geographic areas, this supports the hypothesis that environmental factors acting in utero, such as EDCs, may be responsible.

4.3 Hypospadias and cryptorchidism
Besides declining sperm counts and increasing rates of testicular cancer, the incidence of congenital malformations of the male genital tract is increasing, largely manifest by hypospadias (foreshortened urethra) (13) and cryptorchidism (testis non-descent) (14). However, these observations must be treated with caution since, unlike testicular cancer, the basis of diagnosis is not clear and may have been interpreted differently in various hospitals. In the 1950's Scorer examined over 3500 male babies in a London hospital using very specific definitions of cryptorchidism (86). Using the same diagnosis criteria a more recent Oxford based study of 7500 male births between 1984 and 1988 revealed that the incidence had almost doubled from 0.96 percent to 1.85 percent (87). Studies conducted in England and Wales (13), Sweden, Denmark and Norway (88) show a corresponding increase in the occurrence of hypospadias.

4.4 DES exposed women and their offspring
It is clearly ethically impossible to directly study the in vivo effects of EDCs in humans, and this dictates the need for suitable animal models. However retrospective studies of children from mothers treated with synthetic estrogens during pregnancy has inadvertently provided some human in vitro data. Indeed important insights can be gained into how altering the in utero environment can affect human development.

Diethylstilbestrol (DES), a synthetic estrogen, was prescribed to pregnant women from the late 1940's to prevent miscarriage and pregnancy complications (89). Unfortunately, controlled studies have proved that DES was not effective in treating the disorders for which it was prescribed (90,91), although it continued to be administered until the early 1970's. Several studies since then have indicated serious long-term consequences of in utero exposure to DES for both male (reviewed in (92)) and female offspring (reviewed in (93))

Sons exposed in utero to DES also exhibit several structural and functional abnormalities of the genital tract. Gill (94) reported an increase in the incidence of epididymal cysts and hypoplastic testes in DES exposed males. Further investigation of males with hypoplastic testis revealed a history of cryptorchidism. There is no conclusive evidence linking DES exposure with increased risk of testicular cancer (95), although cryptorchidism is in itself associated with an increasing incidence of testicular cancer. The same study reported a significant decrease in the sperm number in DES-exposed males compared to placebo-exposed men which was confirmed by Stenchever's group (96). Furthermore, it has been shown that men exposed to DES before week 11 of gestation have twice as high a frequency of genital malformations compared to those exposed later (97) illustrating that the impact of DES is dependant on the gestational window of exposure. This is also important in relation to EDC action.

The effects of DES on female progeny have been extensively studied and pathologies include adenosis (98), clear cell adenocarcinoma (99), and structural defects of the cervix, vagina (98), uterus (100), and fallopian tubes (101) which are associated with adverse outcomes of pregnancy and other possible deleterious health problems that are not yet apparent.

The resemblance of abnormalities caused by in utero DES exposure to both observations in the general human population and wildlife studies prompted researchers to determine the cellular and molecular mechanisms of DES using animal models. McLachlan and co-workers have performed a large series of experiments using a murine model of in utero DES exposure using maternal doses of 100 mg/kg on days 9 to 16 of pregnancy (102). Male offspring from these pregnancies displayed retained testes and Mullerian remnants, abnormal sperm morphology and reduced motility, lesions in the reproductive tract, and abnormal reproductive tract secretions (102). Therefore in utero exposure to the synthetic estrogen DES causes reproductive problems in experimental animals that are highly analogous to observations in humans.

Potential mechanisms for DES action have been suggested although these have predominately focussed on the female offspring (103). However, studies in the male have reported abnormal expression of estrogen-regulated lactoferrin in the seminal vesicles of DES-treated male mice (104). The implications of this with respect to the effects of DES on reproduction remain to be determined.

5. Assessing levels of exposure
One major factor to consider when assessing exposure of the fetus to EDCs is bio-accumulation of chemicals in the mother's body. When estimating human exposure to EDCs it is therefore important to consider the diet and lifestyle of individual subjects which will in turn influence total accumulation of EDCs. For example, Inuit populations are exposed to organochlorines through their traditional diet which includes large quantities of sea mammal fat. Researchers found levels of PCBs and dioxin-like compounds were close to those which induced adverse health effects in laboratory animals (105). Furthermore pregnant women from the Disko Bay area in Greenland had high circulating levels of pesticides and PCBs (106) and breast milk samples from Inuit women were reported to have significant levels of contaminants (107). The latter observation reflects the mobilisation of fat during lactation. It follows therefore that children of breast feeding mothers will also be exposed to EDCs during neonatal life. In all these cases the downstream consequences for human male and female offspring remain to be determined. Fat mobilisation, likely associated with EDC release, also occurs during pregnancy hence both the developing fetus and the neonate will be exposed to the mothers chemical legacy stored in fat reserves throughout her life. In animal models of EDC exposure this phenomenon is often overlooked - traditionally test animals are only treated with EDCs for specific windows during gestation. In order to mimic real-life situations authentically, animals should subjected to a ‘lipid bio-accumulation phase’ before pregnancy is established. In this way EDC release by lipid mobilisation during pregnancy is accounted for and not simply exposure from EDCs in the mother’s circulation.

Endogenous hormones (and phytoestrogens) are transported in the blood bound to SHBGs until they reach their target cell whereupon they are released to exert an effect either immediately or following target organ conversion. Consequently, hormone concentrations are tightly regulated in vivo to prevent inappropriate hormone signals (108). However, most EDCs tested to date do not bind to SHBG, therefore they are transported in the maternal blood in an active form. As a result, only low levels of EDCs are required to induce an effect. Furthermore, recent studies show that DDT and DDE are present in human amniotic fluid (109) illustrating that EDCs cross the placental barrier, traditionally thought to protect the fetus from harmful toxins.

Another consideration when testing EDCs is how to mimic bio-accumulation of several EDCs causing a complex "cocktail" effect in the fetus. When tested simultaneously, certain EDCs have been shown to be additive and some researchers have found a synergistic effect (110). While the synergistic issue is contentious, exposure to low levels of several hundred chemicals is more likely than exposure to one chemical at a high dose.

6. Models of EDC actions and effects

6.1 Animal models of EDC effects
While ethical constraints and tissue availability limits studies of testis development and EDC action in the human, much work has been carried out using animal models of in utero EDC exposure.

The male reproductive system is adversely affected by EDCs in several animal species including rats, mice and sheep (Table 3). The rat studies have predominantly focussed on in utero exposure around day 11 and day 15 of pregnancy, corresponding to the development of the fetal testis and the onset of Leydig cell steroidogenesis respectively. Accordingly, a number of deleterious effects have been observed which are linked to hormone mediated events in the gonad.

6.1.1 Pesticide models:
There is an inherent difference between pesticides and other environmental chemicals: pesticides are approved by the authorities for the use in the production of food crops. Therefore, low residual levels of pesticides are allowed in foods and this is closely regulated. Reports attempting to determine the extent of human exposure to pesticides estimate that general intake is approximately less than 0.003 m g/kg/d (111).

Metabolites of DDT, such as p,p'-DDE, and vinclozolin are known to bind predominantly to the androgen receptor (AR) and consequently exert their effects through anti-androgenic behaviour (112). In male rat offspring, exposure to p,p'-DDE (100 mg/kg/day) from gestational day 14-18 caused reduced anogenital distance and retention of thoracic nipples, both strong indicators of anti-androgenic mechanisms (113). Perinatal exposure to vinclozolin produced similar effects accompanied with hypospadias, ectopic testes, vaginal pouch formation and agenesis of the ventral prostate (114). p,p'-DDE have also been shown to act through a mechanism of AR antagonism to alter the expression profile of hepatic cytochrome P450 enzymes involved in testosterone metabolism (Table 3) (19).

The pesticide 1,2-dibromo-3-chloropropane (BDCP) administered to the pregnant rat for 6 days (14.5 - 19.5 days gestation) affects the fetal Leydig cells as demonstrated by a significant decrease in testosterone content in the fetal testis (Table 3) (17). Additionally, the seminiferous tubules were prevented from forming or severely disrupted after fetal exposure during this period (Table 3), although this was not seen in animals treated for a shorter period later in gestation (18.5 - 19.5 days gestation). Some animals in this latter group had testes undergoing normal spermatogenesis while others had tubules comprising only Sertoli cells. This clearly demonstrates the selective effects of EDCs determined by the duration and timing of exposure. These data parallel the gestation-dependant effects of DES in the human (see section 5.4).

Other organochlorine compounds (e.g. methoxychlor) can be metabolised by liver enzymes into compounds with estrogenic activity: mono- and bis-hydroxymethoxychlor (115). Consequently fetal exposure to methoxychlor increased adult mouse prostate size (Table 2) (18) at a dose (20 mg/kg) deemed ‘safe’ for human consumption (below the reference dose) based on traditional ‘high dose’ studies with this chemical. Surprisingly, prostrate enlargement by methoxychlor appeared to be greater than the enlargement produced by in utero exposure to estradiol or DES. A higher dose of methoxychlor (2000 mg/kg) permanently increased both prostate and seminal vesicle weight. Research by Gray and co-workers suggests that the bis-hydroxy metabolite of methoxychlor works through multiple tissue specific mechanisms (116). Hence combined androgen-antagonism and estrogen-agonism of methoxychlor may cause more severe endocrine disruption compared with ‘estrogen-only’ controls.

6.1.2 Dioxin models:
Low levels of dioxins have been detected in many foods, however due to their lipophilic nature, especially high levels have been reported in fatty foods such as cow’s milk. Early in lactation human breast milk also contains significant concentrations of dioxin which decreases in time as the mothers body burden of dioxin is transferred to the new-born (117). Consequently breast milk from mothers nursing their second child are estimated to be lower than for primiparous women (118).

Studies in the rat have illustrated that only 0.1% of the maternal dose of TCDD reaches each fetus transplacentally (119,120), compared to more than 7% of the maternal dose accumulating in each pup following lactational exposure (120). This clearly illustrates the importance of the lactational period for TCDD exposure. More recently, decreases in plasma testosterone, alterations in external male genitalia and reductions in sperm reserves have been observed after maternal dosing with TCDD at 1.0 m g/kg on day 15 of pregnancy (Table 3) (20). Furthermore, in utero exposure alone was sufficient to cause most of the effects observed after combined in utero and lactational exposure (20). Since a single dose given at one gestational time point is unlikely to increase maternal fat levels of TCDD, and since only 0.1% crosses the placenta these findings indicate that the fetus is exceptionally sensitive to small changes in endocrine environment.

6.1.3 Phthalate models:
The ubiquitous distribution of phthalates in various foodstuffs results in an estimated daily total phthalate intake as 4.37 mg per person (121). Sharman et al. (121) reported total phthalates in dairy products ranging from 4 to 20 mg/kg. Allowing for bio-accumulation in body lipid, mobilisation of the fat reserves during pregnancy and the lack of protection normally offered by SHBG, it is possible that the developing human fetus could be exposed to significant levels of phthalate compounds during gestation.

It is already known that phthalates cause deleterious effects such as testicular atrophy (122,123) in the adult rat, however little is known regarding effects of in utero exposure. Several rat models of in utero phthalate exposure have been established (21-23,124-126) with exposure levels ranging from 0.2 mg/kg/d (124) to 1500 mg/kg/d (23). Sperm counts were shown to decrease in rats treated with di(2-ethylhexyl) phthalate (DEHP) during pregnancy at higher doses (mg/kg/d) and this was accompanied with changes in testicular enzyme activity (Table 3) (21).

Disruption of Sertoli cells by phthalates does more than simply disturb testis function. Other workers have shown that phthalates also affect testis descent (Table 3) (22,23). Fetal testes from rats treated in utero with mono-n-butyl phthalate (MBP, 1000 mg/kg/d) on days 15 to 18 of pregnancy were located consistently higher in the abdominal cavity than control animals (22). At birth this resulted in a high incidence of cryptorchidism, similar to observations in estrogen-treated fetal mice (127). Thus it is thought that maternally administered MBP may disrupt Sertoli cells and inhibit the secretion of one or more factors involved in migration of the testis (127). A possible role of MIS in testis decent remains to be clarified (see section 6.2.1). Indeed, more recently di-n-butyl phthalate (DBP) has been shown to cause cryptorchid testes when administered maternally on days 12 to 14 (1500 mg/kg/d) or days 15 to 17 (500-1500 mg/kg/d) of pregnancy (23). Fetal exposure after day 17 of pregnancy however did not result in a significant increase in cryptorchidism at any exposure level. Other experimental studies in rats have indicated that dosing with DBP (500 mg/kg/day) from gestational days 12 to 21 causes several adverse effects in pups including retained nipples, hypospadias and absent or partially developed epididymis, vas deferens, seminal vesicles and ventral prostate (126). Effects in adulthood included decreased testis weight, widespread seminiferous tubule degeneration and interstitial cell hyperplasia and adenoma. This is indicative of an anti-androgenic and not estrogenic mechanism of action, however the anti-androgen flutamide causes a different pattern of anti-androgenic effects (128). The authors concluded that DBP is an example of an anti-androgen that perturbs androgen-dependant male sexual differentiation without interacting directly with the androgen receptor. Both DBP and DEHP have been implicated in an increase in testicular and epididymal lesions in male offspring although the mechanism for this effect is unclear (125).

6.1.4 PCB models:
The levels of PCBs in foods have declined since the 1970s. Indeed it has been estimated that the intake of PCBs from fish consumption in Nordic countries is less than 0.4 m g/kg/d (50). Although the lowest dose used in the animal studies described below was 10 m g/kg, the developing fetus may be exposed to levels approaching this due to bio-magnification in the body lipid.

Different PCB congeners have been shown to have estrogenic or anti-estrogenic effects on MCF-7 human breast cancer cells (53). However little data is available concerning fetal exposure to these congeners. Some information is beginning to appear from studies of laboratory animals but due to insufficient testing it is difficult to draw significant conclusions (24,129).

Anti-estrogenic effects were observed with 3,3',4,4'-tetrachlorobiphenyl (PCB 77), a congener known to bind to the aryl hydrocarbon (Ah) receptor (53). PCB 77 has been tested in a rat model using 100 m g/kg maternal administration at day 15 of pregnancy (24). This dose was sufficient to cause a decrease in serum testosterone and seminal vesicle weight, whereas the testis and brain weight increased accompanied by an increase in daily sperm production (Table 2). The authors suggested these effects could be attributed to neonatal hypothyroidism induced by the substance during early fetal development although this has not been confirmed. The expression of the Ah receptor on placenta suggests that these effects may also be due to reduced placental function/growth (130). Altered hepatic bio-transformation of steroids may also play a role.

In an earlier study maternal administration of Aroclor 1254, a commercial mixture of PCB congeners, induced the expression of P450 isoenymes which hydroxylate and thus deactivate steroid hormones including testosterone (Table 2) (25). This provides a mechanism to explain the decrease in circulating testosterone observed in the PCB 77 study (24).

Studies with 3,3',4,4',5-pentachlorobiphenyl (PCB 126) elicited a different spectrum of effects: while serum testosterone and brain weight followed the same trend as PCB 77 exposure, ventral prostate weight was reduced and the test animals displayed altered sexual behaviour (Table 2) (24). This was not surprising since Perakis & Stylianopoulou showed that de-feminisation of the rat brain occurred in the first week or so after birth (131). PCB 126 therefore must have sufficient potency to alter neuroendocrine pathways during gestation, hence causing latent effects in the neonate. This is in contrast to the dioxin study by Bjerke et al. (20) in which lactational not gestational exposure to TCDD was essential to alter sexual behaviour of male offspring (see section 6.1.2). Hence, chemicals from different EDC groups may act during different developmental windows to elicit similar effects.

6.1.5 Bisphenol A models
Vom Saal and colleagues studied the effect of in utero exposure to bisphenol-A (2-20 m g/kg/d) in mice from day 11 to day 17 of pregnancy (26). The lowest dose was calculated as less than the amount swallowed by patients during the first hour after application of a plastic dental sealant (65). Low levels of exposure (2 m g/kg/day) were associated with significant differences in tissue weights (Table 3). Increasing bisphenol A to 20 m g/kg/day resulted in a decline in the subsequent efficiency of sperm production (daily sperm production per g testis) in the adult (Table 3). However, repetition of these studies has failed to produced similar effects (124) (132): indeed an increase in body weight was the only statistically significant change (124). Furthermore, increasing exposure by a factor of ten still yielded no discernible differences (132). Several reasons are suggested for this discrepancy. These include differences in housing conditions, diet, control animal weights and genetic drift of animals over time.

6.1.6 4-octylphenol(OP) models:
The effects of in utero exposure to OP (2-20 m g/kg/d) in mice from day 11 to day 17 of pregnancy have been investigated. Low exposure (2 m g/kg/d) resulted in significant decreases in efficiency of sperm production and number of sperm produced (Table 3), whereas this was not observed at the higher dose (20 m g/kg/d) (26). These results indicate that it may be impossible to assess endocrine disrupting activities in response to low doses predicted to be within the range of human exposure by using high, toxicological doses of EDCs. In addition, an earlier attempt to assess the effects of OP produced no significant effects in the test animals relative to controls for both doses (133). The reasons for this difference remain to be confirmed. However, further studies have shown that steroidogenic factor 1 (SF-1/Ad4BP) and cytochrome P450 17a -hydroxylase/C17-21 lyase (P450c17) expression in fetal testis was decreased after administration of OP (600 mg/kg) to rats from day 11.5 to 15.5 of pregnancy (Table 3) (27,28). Although the mechanism is uncertain, these studies clearly show that a number of functional genes are affected by in vivo EDC exposure. These changes may therefore be associated with abnormal development and function of the gonads in adult life. However despite the above observations, it is important to note that species differences exist between rodents and primates with respect to the fetal HPT axis.

The effects of OP have also been tested on the development of the fetal sheep HPT axis (134). These surgical in vivo studies showed suppressed fetal FSH concentrations in fetuses from mothers treated with OP from day 110-115 of gestation (134) (Table 3). Furthermore, maternal exposure to OP from day 70 of gestation to birth suppressed FSHb mRNA levels and the number of FSHb-immunopositive cells in the pituitary at birth (134). In addition, the same conditions resulted in reduced testis weight and reduced Sertoli cell number in the testis compared to new-born control lambs (134) (Table 3).

Although these findings should be treated cautiously due to the relatively high level of OP employed (1000 m g/kg/d), virtually nothing is known regarding the actual magnitude of human exposure to these compounds. However, levels of APE breakdown products including OP have been reported to range from nanograms to milligrams per litre in waterways throughout Europe and America (135,136). In addition, bio-accumulation of OP in the body lipid indicates that in heavily contaminated areas concentrations similar to those employed above could be achieved.

6.2 Transgenic lessons
Several transgenic animals exist whose phenotype includes impairment of gonadal development and/or function analogous to that seen in the animal EDC exposure models. It follows therefore that transgenic animals have an important role to play in elucidating the mechanism of action of EDCs. Transgenic or knockout mice facilitate the identification of genes critical for normal male reproductive development and function. Important examples of this are as follows:

A. steroid receptors
Studies of the two estrogen receptor (ER) isoforms (a and b) suggest that the b form of the receptor predominates in the fetal testis. Despite this, male ERb knockouts are fertile whereas ERa knockouts are infertile and have smaller gonads (reviewed in (137)). Male double ER mutants (abERKO) are also infertile but exhibit apparently normal reproductive tract development (138). However, adult abERKO males have reduced epididymal sperm counts and decreased sperm motility (138). The Tfm mouse (Tfm = testicular feminisation) is a naturally occurring androgen resistant mutant. These animals have been used to show that androgens are essential for the second phase of testicular descent. These transgenic mice are therefore useful biological tools in determining different mechanisms of EDC action (i.e. androgenic, estrogenic, anti-androgenic or anti-estrogenic).

B. insulin-like factor 3
The Insl3 gene encodes insulin-like factor 3 (Insl3) and is specifically expressed in Leydig cells of the fetal and postnatal testis. Male homozygous Insl3 knockout mice suffer from bilateral cryptorchidism due to failure of gubernaculum development and are therefore infertile (139). Furthermore, Insl3 mRNA is decreased in cryptorchid DES-exposed fetuses (140) providing a possible mechanism by which synthetic estrogens cause this condition.

C. apoptosis
Apoptosis is an important feature during testis development and this is partially controlled through the expression of a large family of related apoptotic regulatory genes. The prototype apoptosis inhibitor is the protein product of the proto-oncogene bcl-2. The gene product forms homo- or hetero-dimers with apoptosis inducing factors such as bax. Over expression of bax negates the death repressor effects of bcl-2 and promotes apoptotic cell death (141), hence the bcl-2:bax ratio is one important component of cell survival. Indeed mice over-expressing bcl-2 exhibit abnormal adult spermatogenesis (142) and bax-deficient male mice are infertile as a result of disordered seminiferous tubules characterised by an accumulation of atypical pre-meiotic germ cells (143).

The expression of Fas ligand (FasL) and Fas receptor (FasR) are also important in the regulation of apoptosis. When FasL cross links with it’s receptor, apoptosis is induced in FasR-positive cells. Although this system is reported to be important for normal adult spermatogenesis, Fas knockout mice and lpr mice (spontaneous mutant of FasR gene) have normal testicular function. Since FasR is localised to adult rodent germ cells and FasL to the Sertoli cells (144,145) the role of this system in adult testis function remains to be clarified. Interestingly, both testicular FasR and FasL are up-regulated in response to mono-(2-ethylhexyl) phthalate (MEHP) exposure (144) and the implications of this for fertility remain to be determined. In addition, an estrogen response element (ERE) motif has been identified in the promoter region of the FasL gene (146). These findings suggest that EDCs recognised by the estrogen receptor could disrupt FasL expression in the testis, hence affecting testis development and function. However, human studies have failed to detect FasR or FasL mRNA in fetal testis at 20-22 weeks of gestation (147) hence EDC action through this mechanism may be limited to certain gestational windows. Furthermore, it is recognised that many other apoptosis regulatory genes may be important for normal testis function.

6.3 The human dimension
In order to understand the mechanisms by which EDCs affect development of the human fetal gonad, it is important to determine the key processes regulating human gonad development. Extrapolation from animal models is not always possible due to basic differences in control mechanisms (eg. LH-dependant vs. LH-independarnt) and/or developmental timing (eg. testis descent) in the human compared with rodent species. Thus, in utero developmental processes susceptible to EDC perturbation in the human may be different to those in rodent models or may occur at a different gestation. Ethical committee guidelines have limited the collection of human fetal tissues to earlier than 21 weeks of gestation.

6.3.1 Development of the human gonad
Testis development is initiated in the embryo as a response to the expression of the sex-determining gene, SRY. In addition to SRY, there are several downstream effectors and autosomal genes (e.g. SOX-9 and SF-1) that are required for normal differentiation of the male gonad (148,149).

The appearance of Sertoli cells at 6-7 weeks is one of the first morphological signs of testis development. Testicular cords are formed by the aggregation of Sertoli cells and the germ cells are later enclosed within these structures. Between 10 and 15 weeks, Sertoli cells in the newly differentiated testis produce Mullerian inhibiting substance (MIS). This causes regression of the Mullerian ducts, which would otherwise differentiate into the female genitalia. Persistence of these ducts is associated with failed testis descent, hence MIS is thought to be critical for masculinisation and thus a likely target for EDCs. However, normal testis decent in MIS knock out mice suggests that MIS is not crucial to this process (150). Alternative roles of MIS in testis development remain to be clarified.

Sertoli cell multiplication occurs during fetal, neonatal and prepubertal life (151,152) and is dependant on FSH stimulation. Circulating FSH levels have been shown to rise accordingly as gestation progresses from 17 to 40 weeks (153). Inhibin production, which is reported to predominantly occur in the adult Sertoli cells, suppresses FSH production by the pituitary (154). All three inhibin sub-unit mRNAs (a, bB and bA) have been reported in human fetal testis (155-157); while only a and bB sub-unit protein have been detected during gestation (158). Comparison of inhibin B with FSH levels at mid pregnancy found a negative correlation coefficient (159). In the adult testis each Sertoli cell "nurses" a fixed number of germ cells into spermatozoa (160,161). It follows therefore that the establishment of a normal adult sperm count is dependant on the generation of Sertoli cells. Hence, adequate Sertoli cell proliferation during fetal, postnatal and prepubertal windows is vital to establish a normal sperm count in adulthood.

The interstitial fetal Leydig cells begin differentiation during the eighth week of gestation (162) and secrete testosterone to promote differentiation of the Wolffian ducts into the seminal vesicles, vas deferens and epididymis. It seems that the onset of testosterone formation may be independent of gonadotrophin control (163). However, the testosterone-LH negative feedback loop is thought to be functional during the second trimester of pregnancy as pituitary and serum LH concentrations are lower in male than female fetuses during this period (153,164). The Leydig cells undergo rapid proliferation until week 18 when a plateau is reached and Leydig cell number remains at this level until the third trimester (165,166). Studies in our laboratory have confirmed that there are significant changes in the number of proliferating interstitial cells (Figure 1a, P<0.05, figure 1c). This is accompanied by a steady increase in levels of the steroidogenic enzymes 3bHSD and P450c17 (Figure 1b, P<0.01, figure 1e, 1f (167). Although the Leydig population increased in number between 13 and 19 weeks, the proliferating interstitial cells were not steroidogenic (Figure 1d, brown = PCNA, pink = 3bHSD). This indicates that Leydig cell precursors exist in the interstitial region before differentiation and after undergoing a phase of rapid division, they are recruited into the steroid secreting cell population. The period of maximum Leydig cell number is also coincident with a peak in fetal testosterone synthesis between 14 and 18 weeks of gestation (162) equivalent to levels observed in the adult male. Hence, this period of Leydig cell hyperplasia is critical for normal masculinisation of the reproductive tract. In addition, dihydrotestosterone levels in maternal and fetal plasma are stable from mid-pregnancy to term while amniotic fluid concentrations increase by 75%. Furthermore, while there are no sex differences in fetal plasma or amniotic fluid dihydrotestosterone values throughout pregnancy, male testosterone levels are different than the female (168).

In agreement with studies in the rodent, the beta form of the estrogen receptor is found in most cells of human adult testes, including Sertoli cells (169) and developing spermatids (170). In the human fetal testis, levels of ERb are reported to be higher than ERa at the mRNA level. Our own observations have confirmed the expression of both receptor forms at the protein level, however ERa is present in the interstitium and tubule areas throughout gestation (Figure 2a), whereas using a monoclonal antibody to ERb (epitope at human N-terminus) staining is confined to the interstitial area (Figure 2b) (171). In contrast, AR is specifically localised to the peritubular myoid cells (PMCs) immediately surrounding the seminiferous tubules (Figure 2c). This contrasts with studies in adult human testis where Sertoli cells expressed higher levels of AR than the PMCs (172).

6.3.2 EDCs and development of the human testis
EDCs may adversely influence gonad development in a number of ways:

A. Hormone receptors
Since hormone receptors are distributed throughout the human fetal testis, EDCs have several cell types to target. Androgenic or anti-androgenic chemicals may target the fetal peritubular myoid cells. Although the function of these cells is not fully understood, they are thought to be involved in Sertoli/Leydig cell interactions. In contrast estrogenic or anti-estrogenic chemicals may influence the interstitial cells and to some extent the cells within the seminiferous tubules depending on whether the EDC acts through the alpha or beta form of the receptor.

B. Fetal HPT axis
In utero exposure to OP reduces circulating fetal FSH levels (see section 6.1.6).

C. Hepatic biotransformation of steroids
Reduced circulating testosterone levels have been associated with altered expression of steroid metabolising enzymes in the liver (see sections 6.1.1 and 6.1.4).

D. Altered steroidogenic enzyme expression in the placenta and other fetal tissues
Recently, human fetal tissues have shown expression of 17b-hydroxysteroid dehydrogenase isozymes which are capable of regulating the relative concentrations of estrogens and androgens locally (173). The authors speculated that these isozymes may be crucial in the prevention of in utero exposure of the fetus to excessive estradiol from the maternal circulation and amniotic fluids.

E. Altered expression/function of apoptosis regulatory genes
We have recently shown that bcl-2 levels are depressed in human fetal testis incubated with the EDC dieldrin at doses estimated to be within the range of current human exposure (10^-9M) (174). This was accompanied by a significant decrease in testosterone secretion relative to controls (175). In the fetal testis bcl-2 is predominantly localised to the peritubular myoid cells (Figure 2d) which are also positive for AR protein (Figure 2c). Bcl-2 immunopositive cells are also found in the interstitium where ERa and ERb are expressed (Figure 2a, 2b).

This suggests several possible pathways by which dieldrin may affect testicular bcl-2 and hence fetal testis development and function (Figure 3). Dieldrin is estrogenic (176,177), and therefore can bind to the ERa/b-positive Leydig cells in the interstitium. This could reduce testosterone secretion and thus availability for binding to the AR expressed on the PMCs. Since testosterone is reported to protect against testicular apoptosis, this may be associated with reduced levels of bcl-2 in the PMCs resulting in apoptosis (Figure 3b). The further downstream consequences of this are not known but could include reduced androgen receptor-mediated processes.

Dieldrin is also anti-androgenic (178) and may therefore directly target the androgen receptor on the PMCs and block binding of testosterone. As described above, this may lead to increased apoptosis associated with decreased bcl-2 (Figure 3c).

Alternatively dieldrin may not act through a non-receptor-mediated mechanism within the fetal testis. A dieldrin-induced reduction in Leydig cell numbers may cause a decrease in testosterone production with obvious downstream consequences (Figure 3d).

7. Summary and conclusions

There is now considerable evidence that male reproductive function is declining in human and wildlife populations. This is coincident with the increasing use and prevalence of man-made chemicals in the environment over the last fifty years. Certain chemicals have subsequently been shown to disturb the developing fetal endocrine system of laboratory animals in utero. In these experiments, treatment caused similar male reproductive problems in offspring as those already observed in wildlife and human populations. In addition, both the human DES data and rodent studies have shown that there are specific windows of gestation when the developing fetal gonad is highly sensitive to small endocrine changes. Animal in vivo and human in vitro studies have identified EDC sensitive genes. Consequently, hypotheses are being generated concerning mechanism of action e.g. disturbed testicular apoptosis and altered hepatic biotransformation of steroids.

While animal studies provide us with valuable insights into the range of effects that can be attributed to in utero EDC exposure, varying maternal doses employed by different research groups make relation of the results to human observations difficult. The EDC concentration representative of fetal exposure levels is uncertain. Confounding factors include: (a) the vast number of chemicals termed EDCs, (b) the ability of chemicals to bioaccumulate in body lipid, (c) the metabolism of body lipid during pregnancy releasing the mothers lifetime EDC legacy into circulation and (d) the poorly understood kinetics of EDC transfer across the placenta. Thus, the level of fetal exposure can only be crudely estimated at present. This highlights the need for large animal models of EDC in utero exposure where the partitioning of EDCs between the mother and fetus and transfer across the placenta can be studied in detail.

Despite considerable effort the mechanisms by which these endocrine disrupting chemicals exert their effects are still largely unknown. Further studies of the mechanism of action, and consequences, of EDCs in fetal development must be done in order to elucidate how EDCs exert their effects. This can only be achieved using a combined approach whereby animal models are used in combination with in vitro human studies. In conclusion however, there are now sufficient animal model data to prove that EDCs can adversely affect reproductive development and function in the male. Our further understanding of the mechanisms involved may allow intervention strategies whereby we can at least prevent a further decline in male as well as female reproductive health.

For personal use. Only reproduce with permission from SIEP.

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Table 1
Examples of wildlife populations with disrupted sexual development attributable to EDC exposure

Table 2
Differences between man-made EDCs and naturally occurring phytoestrogens

Table 3
Examples of reproductive effects observed in animal models following in utero exposure to EDCs

Figure 1
Schematic representation of the differentiation and subsequent endocrine secretions of human fetal testis.

Proliferation and steroidogenesis in the human fetal testis between 13 and 19 weeks of gestation. (a) Immunolocalisation of PCNA in human fetal testis at 15-16 weeks shows positive interstitial cells (arrows). Numerous PCNA immunopositive Sertoli cells were also noted in the tubules (T). (b) Double immunostaining for PCNA and 3bHSD at 15-16 weeks showed the proliferating interstitial cells (arrows, brown staining) to be immunonegative for 3bHSD (asterisks, pink staining). (c) Quantificaton of PCNA immunostaining revealed a significant peak in interstitial area proliferation between 15-17 weeks (u ), while tubule proliferation (< ) was consistently high. (c) 3bHSD and (d) P450c17 were immunolocalised to the Leydig cells (L) in the interstitium at 15-16 weeks. (f) Image analysis confirmed a steady increase in numbers of cells positive for 3bHSD (¡ ) and P450c17 (= ) in human fetal testis between 13 and 19 weeks of gestation. Scale bar = 50 µm unless otherwise stated. Inset = IgG negative control. Where common superscripts are shown above symbols the values represented are significantly different (ANOVA; P<0.05) for each variable, not between variables. Values expressed as mean ± SEM; n = 4.

Figure 3

Steroid receptors in human fetal testis at 18-19 weeks. (a) ERa immunopositive cells were localised to the interstitial area (IA) and tubules (T). (b) In contrast, ERb immunostaining was confined to the interstitial area (asterisks). (c) AR was expressed specifically by the peritubular myoid cells (arrows) which were also strongly bcl-2 positive (d, arrows) at this stage of gestation (18-19 weeks). Scale bar = 25 µm. Inset = IgG negative control.

Figure 4

Schematic representation of mechanisms by which EDCs may induce dysregulation of apoptosis in the human fetal testis. A. Normal function: Testosterone (T), produced by Leydig cells [1], targets the peritubular myoid cell (PMC) where most androgen receptor (AR) is located [2]. Consequently, this cell functions normally and interacts with Sertoli cells [3]. B. EDC action on estrogen receptor: After crossing the placental ‘barrier’, dieldrin (Diel) reaches the fetal testis [4] and docks with the ER expressed by the Leydig cell [5]. This causes estrogen mediated effects, such as decreasing T output by the cell [6] and hence, the AR on the PMC does not receive adequate T [7]. This causes bcl-2 levels to fall resulting in apoptosis of the PMC [8]. Finally, PMC-Sertoli cell interaction is affected [9]. C. EDC action on androgen receptor: Diel docks with the AR expressed by the PMC [10] thus blocking T binding to the AR on the PMC [11]. Inadequate binding of T causes bcl-2 levels to fall resulting in apoptosis of the PMC [12]. Finally, PMC-Sertoli cell interaction is affected [13]. D. EDC action: non-receptor mediated: Unbound Diel present within the Leydig cell [14] may causes bcl-2 levels to fall resulting in apoptosis of the Leydig cell [15]. Consequently, less T is produced by the Leydig population. 17. Hence, the AR on the PMC does not receive adequate T. 18. This causes bcl-2 levels to fall resulting in apoptosis of the PMC. 19. Finally, PMC-Sertoli cell interaction is affected [19].