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Endocrine Disruptors in Food and Water Linked to Homosexuality and Transgenderism

The Daily Knight

Scientists are continuing to sound the alarm about some common chemicals, including the herbicide atrazine, and link them to changes in reproductive health and development. Endocrine disrupting toxic chemicals have been found to feminize male frogs and cause homosexual behavior. Ashley Ahearn reports on how these substances may be affecting human development and behavior.

"The endocrine system is made up of a series of glands throughout the body that control the hormonal messages that direct development. By imitating natural hormones– such as estrogen and androgen – endocrine disrupting chemicals prevent the body from sending and receiving those messages. Dr. Stephen Rosenthal, a pediatric endocrinologist at the University of California San Francisco, broke down some basic human developmental biology for me. He says in the womb, we all start out developing as girls."

National Library of Medicine


Anthropogenic contaminants in water can impose risks to reproductive health. Most of these compounds are known to be endocrine disrupting chemicals (EDCs). EDCs can impact the endocrine system and subsequently impair the development and fertility of non-human animals and humans. The source of chemical contamination in water is diverse, originating from byproducts formed during water disinfection processes, release from industry and livestock activity, or therapeutic drugs released into sewage. This review discusses the occurrence of EDCs in water such as disinfection byproducts, fluorinated compounds, bisphenol A, phthalates, pesticides, and estrogens, and it outlines their adverse reproductive effects in non-human animals and humans.


Water safety and quality are fundamental to human development and well-being. Besides the pathogenic risk of microbes, several chemical contaminants present in water due to anthropogenic activities can impose risks to human and non-human animal health [1,2]. According to the United States Environmental Protection Agency (USEPA), the definition of contaminant is any physical, chemical, biological, or radiological substance or matter in water. Chemical contaminants are elements or compounds that can be naturally occurring or human-made [3].

The sources of chemical contamination in water are diverse. Chemicals can be present in water through the disinfection processes, chemical release in source water due to industry and livestock activity, and distribution from system components. Non-human animals and humans can be exposed to these compounds by ingesting, inhaling, or dermal contact with contaminated water. Some of the major chemicals that are known as water contaminants are endocrine disrupting chemicals such as disinfection byproducts, fluorinated substances, bisphenols, phthalates, pesticides, and natural and synthetic estrogens. Exposure to these compounds is associated with adverse health and reproductive outcomes in non-human animals and humans; thus, the presence of these chemicals in water has become a public health concern [4,5,6,7,8].

Studies have shown that the contaminants present in water can impair development, fertility, and reproductive function in non-human mammals, humans, and aquatic wild life. For instance, exposure to water disinfection byproducts in drinking water can cause cardiac anomalies in developing rat and porcine embryos [9,10]. Further, exposures to bisphenol A (BPA) and phthalates are known to reduce fertility in mammals by prematurely activating primordial follicles and altering levels of sex-steroid hormones [11,12,13,14,15]. Pesticides have been detected in drinking water sources, and some of these compounds are known reproductive toxicants. For example, exposure to some pesticides is associated with low sperm count and adverse pregnancy outcomes in non-human animals and humans [16,17,18]. Fluorinated substances also can be found in drinking water. Studies have reported that exposure to perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) was responsible for impairing sperm viability and fecundability in non-human mammals and humans [19,20,21,22].

Moreover, water contaminated with synthetic estrogens can cause adverse pregnancy outcomes in non-human animals [23,24,25]. Collectively, these previous studies have shown that chemical contaminants in surface and drinking water worldwide can negatively influence the fertility and reproductive capacity of non-human animals and humans.

This review will discuss the occurrence of chemicals in water and their adverse reproductive effects in non-human mammals, humans, and aquatic life. Specifically, this review will focus on the following categories of chemicals found in water: disinfection byproducts, fluorinated compounds, BPA, phthalates, pesticides, and estrogens.

Water Disinfection Byproducts

The disinfection of drinking water was one of the most important public health achievements in the last century. The treatment of water with disinfectants such as chlorine substantially reduced the incidence of water-borne diseases, and it contributed to increases in life expectancy [26]. However, the reaction between disinfection agents and organic or inorganic matter in source water can form compounds called water disinfection byproducts (DBPs) [27]. The presence of DBPs in drinking water has become a human health concern because epidemiological studies have demonstrated associations between DBP exposure and an increased risk of cancer development and adverse reproductive outcomes [28,29,30,31,32].

Sources of Exposure to DBPs

Several factors can influence the formation of DBPs in drinking water. The presence of organic matter in source water plays a critical role in the formation of these compounds. Organic matter in water mostly consists of molecules such as fulvic, humic, carboxylic, and free amino acids, which are the primary precursors for formation of DBPs [33]. The chemical composition of source water is also an important factor regarding the formation of DBPs. For instance, in areas where the soil and source water are rich in bromine or iodide, the prevalence of brominated or iodinated DBPs tends to be higher than in areas lower in bromine or iodine [34,35]. Generally, increasing temperatures elevate the formation rates of DBPs. In addition, source water with low pH has been associated with high levels of DBPs because the most reactive form of chlorine, hypoclorous acid, is present in high concentrations in water sources with pHs lower than 7.5. Other important elements for DBP formation are the type and concentration of the disinfectant agent used to treat the water. For example, chlorine is known to have the highest potential to form DBPs, especially haloacetic acids, compared to chloramine, chlorine dioxide, or ozone [36].

A significant number of people are exposed to DBPs because of the widespread use of disinfectant agents to treat the water. The most common route of exposure is ingesting treated water, but other potential sources are consumption of food and beverages that were prepared with treated water [27]. Inhalation and dermal absorption also can occur by using showers, bath tubs, swimming pools, or steam rooms [27,36]. To date, more than 700 DBPs have been identified in drinking water [37]; however, only 11 of these compounds are regulated by the USEPA.

The two major classes of DBPs are called trihalomethanes or total trihalomethanes (THMs or TTHMs) and haloacetic acids (HAAs). THMs were the first DBPs identified, and they are the most prevalent in drinking water [37]. Chloroform, bromoform, bromodichloromethane, and chlorodibromomethane are the four THMs that are currently regulated by the USEPA at the maximum contaminant level (MCL) of 0.080 mg/L [38] (Table 1). From 2013 to 2015, the average levels of TTHMs in US drinking water supplies were 0.03 mg/L [39]. HAAs are the second most prevalent DBPs in drinking water. In 1998, the USEPA first regulated the sum of five HAAs (bromoacetic acid, dibromoacetic acid, chloroacetic acid, dichloroacetic acid, and trichloroacetic acid), creating a group called HAA5. In 2016, the USEPA required monitoring for four additional HAAs, encompassing a group called HAA9. The MCL for HAA5 is 0.060 mg/L (Table 1), and levels in drinking water have been reported to be at or below this number [40].

DBPs have also been identified in swimming pool and spa water [41]. The water from these sources changes with the climate, the number and behavior of the users of pools or spas, activities of the swimmers, body fluids such as sweat and urine, as well as environmental contaminants brought into the pool on the skin (sun protectants, lotions) and clothes (bather load) [37,42]. All these components are suitable for reaction with disinfectant agents used to treat the water and can lead to the formation of DBPs. Daiber et al. reported that the total DBP concentrations are higher in water from pools and spas compared to their respective filling waters, which is likely due to the constant availability of disinfectants and organic matter input from swimmers [43]. Besides posing a risk to swimmers because of dermal absorption of DBPs, swimming pools and spas are a concern for public health because volatile DBPs can be trapped in the pool environment, especially in indoor pools, increasing the possible exposure to DBPs via inhalation [41].

Effects of DBPs on the Reproductive System

Non-Human Animals

The toxicological effects of DBPs on developmental and reproductive outcomes have been studied in non-human animals from embryo development to birth. Teixido et al. investigated 10 regulated DBPs (four THMs, five HAAs, and bromate) to assess the developmental toxicity and genotoxicity of these compounds in zebrafish embryos. The authors reported that DBPs caused adverse developmental effects, significant reductions in the tail length (THMs exposure), and increases in malformation rates (tribromoacetic acid, dichloroacetic acid, and bromate exposure) [44]. In a different study, the developmental toxicity of 15 DBPs was assessed using the zebra fish embryo model. The toxicity rank order reported was: acetamides > HAAs > acetonitriles ~ nitrosamines. Furthermore, the study showed that brominated and iodinated DBPs tended to be more toxic than their chlorinated analogues [45]. Wang et al. tested the toxicity of halobenzoquinones, which are an emerging class of DBPs that have been detected in drinking water and swimming pool water [46]. They exposed zebrafish embryos to these compounds and compared the effects of halobenzoquinones to those found in zebrafish embryos exposed to HAAs. They showed that halobenzoquinones induced reactive oxygen species (ROS) generation and inhibited the antioxidative response of cells in developing zebrafish, resulting in death, physical malformations, oxidative DNA damage, and apoptosis. They also determined that the acute toxicity and ROS induction of halobenzoquinones was up to 200 times more potent than those induced by HAAs [46] (Table 2).

Besides causing developmental effects in zebrafish embryos, DBPs have been shown to be toxic to mouse, rat, and porcine embryos. In a study using CD-1 mouse embryos, the effects of exposure to different HAAs during a period of 24 h were assessed. Exposure to HAAs resulted in dysmorphogenesis, alterations in development of the neural tube and optic nerves, and abnormal heart development [47]. Andrews et al. exposed rat embryos to various concentrations of dichloro, dibromo, and bromochloroacetic acid (HAAs) for 48 h and then assessed dysmorphology. The primary effects of HAAs observed were dysmorphogenesis, heart defects, and to a lesser extent, prosencephalic, visceral arch, and eye defects. The developmental effect scores for embryos exposed to the combination of HAAs were higher when compared to the effect scores for embryos exposed to the single compounds, suggesting that the developmental toxicity of these DBPs was additive [9]. Further, exposure to environmentally relevant concentrations of bromodichloromethane, a type of THM, caused transcriptomic and epigenomic adaptive modifications compatible with the cardiac anomalies in porcine blastocysts [10] (Table 2).

DBPs also have been shown to disrupt ovarian function, spermatogenesis, and fertility outcomes. To evaluate the effects of dibromoacetic acid on ovarian function, Bodensteiner et al. exposed female Dutch-belted rabbits daily to dibromoacetic acid through drinking water (0, 1, 5, or 50 mg DBA/kg body weight) from gestation day 15 throughout life [48]. They observed that dibromoacetic acid reduced the number of primordial follicles and total healthy follicles in prepubertal rabbits. In adult rabbits, dibromoacetic acid decreased the number of primordial follicles compared to the non-exposed rabbits [48]. In mice, iodoacetic acid inhibited antral follicle growth and reduced estradiol production by ovarian follicles in vitro [49]. To determine the mechanisms by which iodoacetic acid caused these alterations, Gonsioroski et al. [50] analyzed the gene expression and sex steroid hormone levels of mouse ovarian follicles in vitro. They showed that iodoacetic acid dysregulated the expression of apoptotic factors, cell cycle regulators, steroidogenic factors, and estrogen receptors, subsequently disrupting cell proliferation and steroidogenesis [50]. Narotsky et al. assessed the combined toxicity of regulated DBPs (TTHMs, HAAs, or TTHMs and HAAs) on the fertility indices of rats [51]. They observed that all three mixtures caused pregnancy loss and that HAAs alone or HAAs plus TTHMs increased resorption rates. In another study, the reproductive effects of an environmentally relevant mixture of DBPs representative of chlorinated drinking water were evaluated in rats in a multigenerational bioassay. The authors did not observe adverse effects of DBP exposure on pup weight, prenatal loss, pregnancy rate, gestation length, puberty onset in males, growth, estrous cycles, and hormone levels. However, the DBPs delayed puberty for F1 females, reduced caput epidydimal sperm counts in F1 adult males, and increased the incidence of thyroid follicular cell hypertrophy in adult females [52]. In male rats, dibromoacetic acid caused histopathologic changes in the testis and epididymis. Specifically, dibromoacetic acid caused the retention of spermatids, fusion of mature spermatids, and presence of atypical residual bodies in the epithelium and lumen of seminiferous tubules. In addition, the exposure caused distorted sperm heads, vacuolation of the Sertoli cell cytoplasm, vesiculation of the acrosomes of late spermatids, and marked atrophy of the seminiferous tubules [30]. Melnick et al. described similar testicular lesions in mice exposed to dibromoacetic acid. Specifically, lesions were characterized as spermatid retention and large atypical residual bodies in seminiferous tubules, which were suggested to be a result of the impaired degradative function in Sertoli cells [53] (Table 2).


DBPs have been shown to be associated with adverse reproductive outcomes in women and men. For instance, in a retrospective cohort study conducted in Nova Scotia, Canada, consisting of 49,842 women who had a singleton birth between 1988 and 1995, exposure to chloroform and bromodichloromethane were associated with neural tube defects, cardiovascular defects, cleft defects, as well as chromosomal abnormalities [54]. For neural tube defects, the risk was increased with high exposure to bromodichloromethane but not chloroform. Further, a stronger relation between chloroform and chromosomal abnormalities was observed than between bromodichloromethane and chromosomal abnormalities [54]. In another study, Levallois et al. evaluated the association between maternal exposure to DBPs and the risk of delivering a small for-gestational-age neonate. HAA concentrations above the fourth quartile and THM or HAA concentrations above current water standards increased the risk for small for gestational age neonates [55]. In addition, in a study of 7438 singleton term babies in Bradford, England, TTHM exposure during pregnancy was associated with reduced birth weight [56]. Moreover, in a study of 2460 stillbirth cases from 1997 to 2004 in Massachusetts, chloroform and dichloroacetic acid exposures were associated with stillbirths [57]. In China, exposure to TTHMs was associated with decreased sperm concentration and serum testosterone in men [58]. Further, studies found that a GSTT1 polymorphism modified the association between exposure to bromo-THMs and decreased sperm motility. In addition, cytochrome P450 2E1 (CYP2E1) polymorphisms were associated with the internal blood concentrations of chloroform and TTHM [59] (Table 2).

Null Studies

Although some studies show that DBPs are associated with adverse reproductive outcomes, other studies have not found associations. For example, Cummings and Hedge did not observe effects of dibromoacetic acid in drinking water on the number of implantation sites found on gestational day 9, the number of pups per litter, the number of resorptions, or mean pup weight in rats [60]. Further, Weber et al. did not observe the effects of prenatal dibromoacetic acid exposure on daily sperm production, testicular sperm counts, epididymal sperm reserves, the morphology of seminiferous epithelium, or ovarian follicle counts in mice [61]. Narotsky et al. did not observe effects of a mixture of regulated DBPs on fertility, pregnancy maintenance, prenatal survival, postnatal survival, or birth weights in the parental, F1, and F2 generation of rats [62]. In human studies, no associations were found between exposure to DBPs and time to pregnancy, duration of gestation, small size for gestational age, stillbirths, preterm births, or birth weight [63,64,65,66,67,68]. Furthermore, some studies show that poor semen quality is not associated with exposure to DBPs in men [69,70,71].

These inconsistencies in the literature may be due to several factors. In experiments that use non-human animal models, the levels of DBP exposure are not always environmentally relevant, which can lead to discrepant findings. Further, the methods applied to treat non-human animals with DBPs do not always follow the routes of exposure for human and non-human animals (for example gavage versus drinking water). Thus, it is important for future studies to analyze the effects of single DBPs or mixtures of DBPs at environmentally relevant levels using relevant routes of exposure. In human epidemiological studies, differences in the size and genetic variability of the populations and variations in exposure levels makes comparison of results difficult among studies. The incorporation of subject behaviors into exposure evaluation, such as showering and swimming activities or the consumption of bottled or filtered water, could provide a better understanding of individual exposure to DBPs. Finally, few studies have been done on emerging DBPs and the underlying mechanisms of action of DBPs, opening up areas for additional research.

Perfluoroalkyl and Polyfluoroalkyl Substances

Fluorinated substances are a wide group of organic and inorganic substances that contain at least one fluorine atom. A subset of these substances contains carbon atoms, on which all the hydrogen substituents have been replaced by fluorine atoms. These compounds are called perfluoroalkyl and polyfluoroalkyl substances (PFAS) [72]. In perfluoroalkyl substances, all carbons except the last one are attached to fluorines, and the last carbon attaches to the functional group. In polyfluoroalkyl substances, at least one, but not all carbons are attached to fluorines [73]. PFAS are human-made chemicals that have important properties such as hydrophobic and lipophobic nature, and chemical and biological stability. As a result of these properties, PFAS are used in a wide variety of consumer products and are highly persistent in the environment [74]. The presence of these chemicals in the environment is a concern for public health because exposure to PFAS has been associated with an increased incidence of tumors, endocrine disruption, impaired neurodevelopment, and adverse reproductive outcomes in humans and non-human animals [75,76,77,78,79,80,81,82,83,84].

Sources of Exposure to PFAS

According to the USEPA, PFAS can be found in food packaged in PFAS-containing materials, processed with equipment that used PFAS, or grown in PFAS-contaminated soil or water. These compounds also can be found in commercial household products including stain- and water-repellent fabrics, nonstick products (pans), polishes, waxes, paints, cleaning products, and fire-fighting foams (a major source of groundwater contamination at airports and military bases where firefighting training occurs). Moreover, PFAS can be found in the workplace, including production facilities or industries that use these compounds. PFAS also can be found in drinking water, which is typically localized and associated with a specific facility (e.g., manufacturer, landfill, wastewater treatment plant, firefighter training facility). PFAS also can be present in living organisms, including fish, non-human mammals, and humans, where these chemicals have the ability to build up and persist over time [85]. Common PFAS are listed in Table 3.

As a result of the widespread use of PFAS, these chemicals can be found in surface and groundwater and subsequently in drinking water [86,87]. Studies have described the presence of PFAS in tap water in several countries, and the levels of these chemicals can vary largely depending on the location. In a study in Canada, average concentrations of PFOS and PFOA from the Great Lakes area were 3.4 ng/L and 1.8 ng/L, respectively, whereas samples from the rest of Canada had average concentrations of 0.4 and 0.7 ng/L, respectively [86]. In Brazil, the average levels of PFOS and PFOA in tap water were 6.7 ng/L and 2.7 ng/L, whereas in China, they were 3.9 ng/L and 10 ng/L, respectively [88,89]. In tap water samples from the United States (Ohio and Northern Kentucky), the average concentrations of PFOS and PFOA were 7.6 ng/L and 10 ng/L, respectively. PFAS have longer half-lives in humans than non-human animals, suggesting that humans could be more susceptible to PFAS toxicity than non-human animals [90,91,92]. To date, the USEPA does not have MCLs for PFAS in drinking waters, but this agency is analyzing the necessity of creating MCLs for PFOA and PFOS specifically [93].

Effects of PFAS on the Reproductive System

Non-Human Animals

PFAS are known to disrupt reproductive function in non-human animals. Specifically, PFOA exposure damaged seminiferous tubules, increased spermatogonial apoptosis, and decreased testosterone levels in the testes of mice [19]. Exposure to PFOA decreased the number of mated and pregnant females per male mouse and disrupted blood testis barrier integrity [94]. Further, prenatal exposure to PFOA reduced the number of offspring, caused damage in the testes, disrupted reproductive hormones levels, and reduced expression of the Dlk1-Dio3 imprinted cluster in testes in mice [95]. Prenatal exposure to PFOS decreased sperm count and serum testosterone concentration in male rat offspring [20]. Li et al. demonstrated that rats exposed to PFOS during puberty presented delayed Leydig cell maturation, decreased androgen production, reduced expression of cytochrome P450 11A1 (Cyp11a1), cytochrome P450 17A1 (Cyp17a1), and hydroxysteroid 17-Beta dehydrogenase 3 (Hsd17b3), and they disrupted the expression of apoptotic-related genes BCL2 associated X (Bax) and BCL2 apoptosis regulator (Bcl-2) in Leydig cells [96]. In female mice, PFOA exposure caused a delayed or absence of vaginal opening, lack of estrous cycling, decreased ovarian levels of steroidogenic acute regulatory protein (STAR), CYP11A1, 3-Beta dehydrogenase 1 (HSD3B1), and HSD17B1, and reduced protein levels of amphiregulin and hepatocyte growth factor in the mammary glands [97]. In mice, Chen et al. showed that maternal exposure to PFOA inhibited corpus luteum function, decreased levels of serum progesterone, decreased the ovarian expression of Star, Cyp11a1, and Hsd3b1, increased the ovarian expression of tumor protein (p53) and Bax, and reduced the expression of Bcl-2 in the ovary, leading to embryo resorption, reduced fetal growth, and reduced postnatal survival [98]. Furthermore, PFOA exposure induced apoptosis and necrosis in mouse oocytes, which is likely related to reactive oxygen species (ROS) generation and gap junction intercellular communication disruption between the oocyte and the granulosa cells [99]. Working with female rats, Du et al. found that neonatal and juvenile exposure to PFOA or PFOS dysregulated the hypothalamic–pituitary–gonadal (HPG) axis, leading to advanced puberty onset, increased levels of serum luteinizing hormone and estradiol, and the reduced expression of kisspeptin 1 (Kiss1), kisspeptin 1 receptor (Kiss1r), and estrogen receptor alpha (Esr1) in the hypothalamic anteroventral periventricular and arcuate nuclei [81] (Table 4).


PFAS have been associated with reproductive and fertility dysfunction in men and women. In an epidemiologic study in Denmark, men with high combined semen levels of PFOS and PFOA had decreased normal sperm numbers compared to men with low semen levels of PFOS and PFOA [100]. In another study in Denmark, in utero exposure to PFOA was associated with lower sperm concentration and total sperm count in adult men [101]. Further, in vitro exposure to PFOA impaired human sperm penetration in synthetic mucus, which was likely caused by excessive ROS production, compromising human sperm penetration ability and acrosome reaction by canceling progesterone-induced Ca2+ signaling [102]. In addition, men exposed to PFOA for up to 2 h exhibited altered sperm motility due to plasma-membrane disruption [21]. In China, maternal exposure to PFAS was associated with shorter anogenital distance in boys, providing evidence that PFAS may function as EDCs to affect male genital development [110]. Moreover, PFOA and PFOS exposure were associated with reduced semen quality, testicular volume, penile length, and anogenital distance in men in the Veneto region, Italy. This same study demonstrated that PFOA plays an antagonistic role on the binding of testosterone to androgen receptor, possibly dysregulating the HPG axis [112] (Table 4).

In women, exposure to PFAS has been associated with endometriosis in the US and China [103,104,105]. Further, levels of PFAS in blood have been associated with decreased serum levels of estradiol, progesterone, sex hormone-binding globulin, follicle-stimulating hormone (FSH), and testosterone [106,107]. Plasma concentrations of PFAS in pregnant women in the Danish National Birth Cohort were associated with low birth weight and long time to pregnancy [22,108]. Moreover, the maternal–infant research on environmental chemicals study, a cohort study of pregnant women across Canada, showed that plasma levels of PFOA and PFHxS were associated with reduced fecundability [109]. In Swedish women, prenatal exposure to PFOA was associated with higher odds for small for gestational age [113]. In a recent study, PFAS exposure was associated with increased age at menarche and irregular menstrual periods in young women. The same study reported a significant alteration in the expression of genes related to embryo implantation in Ishikawa cells exposed to PFOA compared to non-exposed cells [111] (Table 4).

Although several studies have shown that PFAS exposure causes adverse reproductive and health effects, little is known about emerging PFAS and their effects on the environment and human health. For example, perfluoro-2-propoxypropanoic acid (PFECA), a PFOA replacement known as “GenX”, has been shown to have higher toxicity than PFOA when correcting for differences in toxicokinetics. However, the effects of “GenX” on reproductive outcomes are unclear [114]. In addition, few studies have examined the effects of exposure to a mixture of PFAS on non-human animal and human health, which could provide more information about the potential interactions between individual PFAS. Future studies should include these factors to improve our understanding of PFAS toxicity and adverse health outcomes.

Disphenol A

Bisphenol A (BPA) is an important compound in the bisphenol (bishydroxyarylalkanes) group [115]. Currently, BPA is a high production volume chemical that is widely used in manufacturing polycarbonate plastics and epoxy resins for industrial use [116]. Polycarbonate plastics are used in food and drink packaging (water and infant bottles, compact discs, impact-resistant safety equipment, medical devices), whereas epoxy resins are used as lacquers to coat metal products (food cans, bottle tops, water supply pipes) [117]. Human exposure to BPA is a public health concern because BPA has the ability to bind membrane and nuclear receptors such as androgen, estrogen, and thyroid receptors, causing endocrine disruption, tumors, adverse reproductive outcomes, and transgenerational effects [118,119,120,121,122].

Sources of Exposure to BPA

The primary source of exposure to BPA is diet, but BPA is ubiquitous in the environment, air, dust, and water. BPA can leach into food from the protective internal epoxy resin coatings of canned foods and from consumer products such as polycarbonate tableware, food storage containers, water bottles, and baby bottles [117]. Canada was the first country to prohibit the sale and importation of BPA-containing baby bottles [123]. Several states in the US banned the use of BPA in cups, bottles, thermoses, baby food and infant formula containers, or thermal paper [124]. Further, the French National Assembly and Senate suspended the use of BPA in all applications that have contact with food [125]. In contrast, the European Food Safety Authority concluded that BPA was not a threat for the health of consumers of any age. In addition, the United States Food and Drug Administration (USFDA) declared that BPA is safe at the current levels occurring in foods [126]. Although controversies about BPA regulation exist, studies have shown that this chemical is an endocrine disruptor, which means that this compound is able to trigger adverse health effects at low and environmentally relevant doses [127,128].

BPA is ubiquitous in aquatic environments and can be detected in rivers, effluent from sewage treatment plants, and water from water treatment plants [129,130]. Specifically, the mean concentrations of BPA in the Huangpu River in China were 22.93 ng/L in surface waters, 84.11 ng/g in suspended solids, and 7.13 ng/g dry weight in surface sediments [130]. Further, a study in Taiwan determined that BPA concentrations in drinking water were increased with contact time in polyvinyl chloride (PVC) pipes [131]. In some provinces of South Africa, BPA was found to be present in 62% of the analyzed drinking water and wastewater samples [132]. Further, in raw water and tap water samples in France, BPA levels were up to 1430 ng/L and between 9 and 50 ng/L, respectively [133]. In wastewater treatment plants, BPA was found at concentrations of 60.5 ng/L in five states in India, 1960 ng/L in 49 samples from Xiamen City in China, and 412 ng/L in one sample from Dalian City, China [134,135,136]. The USEPA reported that BPA concentrations in US drinking water are typically below 1 µg/L [137]. Although exposure to BPA through tap water is a minor source of human BPA exposure, bottled mineral water may also lead to exposure [138].

Effects of BPA on the Reproductive System

Non-Human Animals

BPA is known to cause adverse reproductive outcomes in non-human animals. Specifically, it has been demonstrated that BPA disrupts the HPG axis in mice, rats, and zebrafish [139,140,141,142,143,144]. In mice, studies have shown that BPA exposure reduced sperm motility, reduced normal sperm morphology, decreased sperm membrane integrity, decreased sperm count, impaired sperm function, induced spermatocyte apoptosis, and impacted testicular development [145,146,147,148,149]. In females, BPA exposure is known to cause altered mammary gland development and morphology. Specifically, in utero exposure to BPA resulted in altered development, increased epithelial volume, and the altered ductal morphology of mammary glands in mice [121,150]. Further, Ibrahim et al. showed that adult BPA exposure increased the number of the ducts and acini of the mammary gland, with hyperplasia in their lining epithelium in rats [151]. These studies agree that mammary gland changes due to BPA could lead eventually to an increased incidence of mammary gland cancer [119] (Table 5).

Exposure to BPA has been shown to affect the ovaries. Prenatal BPA exposure inhibited germ cell nest breakdown in ovaries of the F1 generation in mice, decreased the num