Water and soil pollution as determinant of water and food quality/contamination and its impact on male fertility

There are numerous substances with a recognized anti-androgenic effect, from air and ground pollutants to plasticizers. In the latter category, phthalates (Ps) are the most investigated compounds as they are employed in virtually all industrial applications and consumer products as additives. Ps are inexpensive synthetic chemicals and have been widely used as plasticizers in a broad range of industrial and commercial products [16, 17]. The most commonly used phthalates are di-(2-ethylhexyl) phthalate (DEHP), di-n-butyl phthalate (DBP), diethyl phthalate (DEP), and benzylbutyl phthalate (BzBP). More than 75% of DEHP produced worldwide is used in plastic products. The other Ps are largely used in personal care products like foams, shampoos, dyes, lubricants, and food packaging materials [18]. Since these compounds are not covalently-bound polymers, their exposure to heat over time has the potential to favour their migration into food [19]. Human exposure to environmental pollutants from foodstuff poses health risk for the general population. Plasticizers such as phthalate esters, because of their anti-androgen and estrogen-like activity, are indicated as major endocrine disruptors. Both in vitro and in vivo toxicology studies have demonstrated their endocrine disrupting potential in model organisms, with endpoints such as antiandrogen effects, reproductive abnormalities, testicular lesions and reduced sperm production [20] (Fig. 1). However, dose ranges used for traditional reproductive toxicological studies were much higher than those observed in human epidemiological studies. Therefore, it is not surprising that these studies do not entirely align with the human studies. Controversially, in vitro and in vivo toxicology studies with low exposures to Ps were linked to decreased semen quality and male infertility in animals, as well as to decreased androgen production and steroidogenesis [21‐30]. Ps have mostly shown the antiandrogen effect on testicular function during steroid formation [31‐33]. Several in vitro or in vivo studies also showed that Ps, as well as their metabolites (e.g., DEHP/MEHP, DBP/ MBP) have stimulatory effects at low doses through inducing the production of progesterone, testosterone, steroidogenesis-related proteins and gene expression [23, 24, 26, 27, 29, 30]. The adverse effects of Ps on sperm quality were confirmed by ex vivo studies, where spermatozoa were incubated in vitro and exposed with high concentrations of phthalates [34]. It was reported that the sperm motility was decreased and that cytotoxicity was caused at long-term exposures (> 3 days) of human semen samples to the metabolite DEHP [34]. In parallel DHEP has been shown to inhibit testosterone production, when cultured in vitro with explants derived from human testes [35].

Epidemiological studies report an association between Ps exposure and altered seminal parameters have been reported [36](Table 1). It is important to note that exposure of infants to Ps is mainly due to both maternal exposure and breastfeeding. In fact, breastmilk levels of the phthalate metabolites are positively associated with maternal diet and water consumption. In Korea, breast feed infants exceeded the reference daily dose of DEHP by 8% and of DBP by 6% [37]. More recently, the urinary levels of Ps metabolites were related to infertile biomarkers and infertility in Chinese men [38].

Studies that were conducted in human populations corroborated the in vitro findings and suggested that exposure to phthalate metabolites is correlated with lower motility of spermatozoa in men from subfertile couples [39]. The DNA damage induced in spermatozoa, the sperm motility and the morphology of the spermatozoa were weakly associated with the exposure to Ps [40‐43], whereas with regard to the disruption of the hormonal function, an inverse association between MEHP exposure and testosterone and oestradiol levels was reported [44].

Data available on the effect of Ps on male reproductive health is limited, largely confined to specific cases of infertility [45]. Ps are rapidly metabolized and excreted in urine and feces and therefore the assessment of exposure to Ps in human relies on the measurement of urinary concentrations of phthalate metabolites. However, little or even no attention is given to the possible accumulation of un-metabolized Ps in different tissues [46]. This evidence rises some concerns about the appropriateness of parameters employed as index of exposure to contaminants, in particular for those substances like Ps that, showing specific tissue accumulation, may exert risk associated to long term exposures [32]. To this regard, quantification of both parent compound and corresponding metabolites in specific body fluids may represent an informative parameter with better correlation with clinical parameters [33].

In addition to phthalates, human exposure to the ED Bisphenol A (BPA) affects endocrine-reproductive function in males (Fig. 1). BPA is a high production-volume chemical that is widely used in the manufacture of consumer products such as polycarbonate plastics, epoxy resin liners of canned foods, some dental sealants and composites, and thermal receipts [47]. Due to its widespread use in consumer products, exposure to BPA is ubiquitous. In the United States, more than 90% of urine samples obtained from participants in the 2003–2004 and 2011–2012 National Health and Nutrition Examination Survey (NHANES) had BPA concentrations above the limit of detection [48, 49]. Exposure to BPA has garnered concern and regulatory attention over the past decade owing to its potential endocrine disrupting effects. Specifically, in vitro studies have shown that BPA binds to ERα and ERβ, producing weak estrogenic activity [50, 51]. BPA has also been cited for its ability to bind to the AR, peroxisome proliferator–activated receptor γ, and thyroid hormone receptor in experimental animal studies [52]. For example, doses below the present lowest observed adverse effect level (LOAEL; < 50 mg/kg) for BPA were associated with decreased sperm counts [53‐57], impaired sperm motility [53, 55, 56, 58], and increased sperm DNA damage [55, 58‐65]. In addition, doses below the present LOAEL for BPA were related to decreased testosterone levels [56, 63, 66‐68]. Most animal studies concluded that BPA was a testicular toxicant [69‐72]. There are differences across studies related to methodologic aspects such as dose, exposure route, timing, and outcomes measured (reviewed in [73]).

In humans, there is a growing body of literature exploring the associations between male urinary BPA concentrations and semen quality parameters, DNA damage, and reproductive hormones [5, 74‐80] (Table 1) and a few studies on paternal urinary BPA concentrations and markers of couple fecundity and fertility such as time to pregnancy and live birth [28, 81]. Only six studies have explored the relationship between urinary BPA concentrations and semen parameters, and two of these studies also examined the association with sperm DNA damage [74, 78]. In the only prospective study to date, Li et al. explored the association of urinary BPA concentrations on semen parameters among 218 factory workers from four regions in China [76]. Their study found a negative association between urinary BPA concentrations and sperm concentration, total sperm count, sperm vitality, and sperm motility. The epidemiologic literature investigating the endocrine disrupting effects of BPA on male reproductive hormones also is limited and presents heterogeneous results. To date, one study has explored this association among men occupationally exposed to BPA [79], two studied the association among men from the general population [77, 80], and two studies investigated this association among either fertile men or subfertile men from a fertility clinic [74, 75]. The association of male urinary BPA concentrations with couple reproductive outcomes was recently assessed in two studies. Using the EARTH study cohort consisting of subfertile couples undergoing fertility treatment at MGH, Dodge et al. examined the associations of paternal urinary BPA concentrations with fertilization, embryo quality, implantation, and live birth among 218 couples who underwent 195 intrauterine inseminations and 211 in vitro fertilization cycles [81]. No associations between paternal urinary BPA concentrations and reproductive outcomes following fertility treatment were found. The association of paternal urinary BPA concentrations with couple reproductive outcomes was also investigated in the LIFE study of 501 couples discontinuing contraception with the intention of becoming pregnant. Similarly to the study among fertility clinic patients, Buck-Louis et al. did not find association between paternal urinary BPA concentrations and time to pregnancy [28], but interestingly higher paternal urinary BPA concentrations were significantly associated with fewer male births. Although the epidemiologic literature on this topic is growing, the evidence supporting an association between urinary BPA concentrations and male reproductive health in humans remains limited and inconclusive (Table 1). Several methodologic differences could explain discrepancies between human studies. First, studies included different study populations, including fertile males who may be less susceptible to the effects of BPA than would sub-fertile men. Second, the distribution of urinary BPA concentrations varied across studies. If there is a nonlinear association between BPA exposure and markers of reproductive health then we may not find consistent results across study populations with markedly different exposure levels. However, it is worth noting that contradictory results were found even among populations with similar urinary concentrations. Moreover, if exposure to BPA is not constant (within-individual variability is known to exist), the time window of BPA exposure captured in cross-sectional studies (e.g., the last 24 h) may not be the biologically relevant exposure window (e.g., the last 90 days for spermatogenesis). Third, all studies measured adult exposure, but none considered early life exposure (e.g., prenatal or peripubertal windows) which may be more sensitive to effects of BPA. Finally, residual confounding factors correlated with BPA exposure and semen quality were not accounted for.

Organophosphate (OPs) pesticides are one of the most widely used class of pesticides for agricultural purposes [82]. They are metabolized by xenobiotic metabolizing enzymes, notably the cytochrome P450 (CYP) and the Paraoxonase (PON) families of enzymes and are therefore not persistent in the environment [83]. The exposure to OPs is assessed by the detection of their corresponding secondary metabolites notably the dialkyl phosphates in biological matrices such as urine [83‐85]. The exposure of humans to OPs can be either professional (chemical plant workers, agricultural workers) or environmental (through soil and water contamination). Organochlorine compounds such as DDT and dioxins are not metabolized by the human body and accumulate for a long period of time [8]. In addition, such compounds appear to be a lot more persistent in the environment compared to organophosphorus compounds. Agonistic effects of MTX, an organochlorine pesticide used as an insecticide that was indented to replace DDT, have been reported for the estrogen receptor subtypes ERα and ERβ, whereas an opposite response was noted for the AR [86‐88]. Thiophosphates, a class of organophosphorous pesticides, inhibit P450 enzymes namely, CYP3A4 and CYP1A2 that are involved in the metabolism of estrone and testosterone in the liver [89]. At the molecular level, EDs can affect the expression of steroid and sex hormone related enzymes by inducing their corresponding transcription, via binding to nuclear receptors. Notably, OPs and dioxins have been documented to bind with considerable potency to the aryl hydrocarbon receptor (AhR) that induces the expression of CYP1 genes that in turn metabolizes estradiol (E2) to hydroxylated derivatives [90]. Specific adverse effects on the male reproductive system have been reported to occur by organochlorine pesticides, such as endosulfan and DDT due to the disruption of the hypothalamic–pituitary testes axis and the direct interaction with the sex steroid receptors in the target tissue [91] (Fig. 1). The occupational exposure to pesticides increases the risk of morphological abnormalities in the sperm of farm workers that includes a decline in sperm count and a decreased percentage of viable sperms. OP pesticides such as parathion and methyl parathion can decrease the concentration of the sperm by damaging the seminiferous epithelium, while it has been suggested that pesticide exposure affects sex accessory glands that may also reduce the seminal volume [92, 93]. The exposure to pesticides reduces the seminal volume, increases the seminal pH and increases the abnormal sperm head morphology [93]. In addition, Lifeng et al. demonstrated that sperm motility could be affected by a limited number of pyrethroid pesticides, such as fenvalerate [94]. However, most studies were cross-sectional and due to different participation rates and lack of information on time dimension of the cause–effect relationship cannot be supported. Further controlled studies are needed to make sure about the effects of pesticides on male infertility. Although animal studies confirmed an impact of these chemicals on reproductive health, it should be noted that rats are more sensitive to the effects of pesticide exposure in comparison to humans. Moreover, synergistic, and potentiating effects of multiple chemicals are rarely explored in toxicological research. In spite of several studies and laboratory researches, no consistent view exists on the role of chronic pesticide exposure on semen parameters at present [91].

Perfluoroalkyl compounds (PFCs) are a class of organic molecules characterized by fluorinated hydrocarbon chains extensively used in industry and consumer products including oil and water repellents, coatings for cookware, carpets and textiles. PFCs possess unique physical chemical properties due to their amphiphilic structures and their strong carbonfluorine bonds. Therefore, long chain PFCs are non-biodegradable and bioaccumulate in the environment [95, 96]. PFCs have been found in humans and in the global environment and their toxicity, environmental fate, and sources of human exposure have been a major subject of research. Currently 23 PFCs are available, which includes perfluorooctanoic acid (PFOA) and perfluooctane sulfonate (PFOS), which are the predominant forms in human and environmental samples. However, the stability that makes PFCs desirable for commercial use, also entails that they are environmental contaminants due to their resistance to various modes of degradation [97].

Both in vitro and animal studies on PFCs toxicity have shown a detrimental effect of PFOA and PFOS on testicular function, by the alteration of steroidogenic machinery and subsequent defect of spermatogenesis [98‐102]. Among the endocrine effects of PFOS in particular, it should be emphasized that this compound can affect the hypothalamic–pituitary axis activity [103, 104] (Fig. 1). It is also able to exert its toxicity at testicular level [105], as reported in rats [103, 106] and in testis models [107]. According to a recent study on male rats [108], high doses of PFOS orally administered for 28 days seem to modify the relative gene and protein receptor expressions of several hormones of the reproductive axis (GnRH, LH, FSH and testosterone) (Fig. 1).

Various PFCs compounds have been found in human serum [109], seminal fluid [110], breast milk [111] and even umbilical cord [112], suggesting a life-long exposure to PFCs in humans, from foetal stages until the adult life. In addition to their persistence, PFOA and PFOS have been shown to induce severe health consequences, such as neonatal mortality, neurotoxicity and immunotoxicity: PFCs act as endocrine disruptors on the foetus and newborns, leading to developmental defects [113]. This has led to strict regulation of PFOA and PFOS use in industrial processes, as the compounds were added to the Annex B of the Stockholm Convention on Persistent Organic Pollutants. In addition to health concerns on the impact of these substances on foetal development, epidemiological studies have focused also on the relationship between PFCs and human fertility, but recent research has focused mostly on female fecundity. In humans, in utero exposure to PFOA was associated later in adult life with lower sperm concentration and total sperm count and with higher levels of luteinizing hormone and follicle-stimulating hormone [114] (Table 1).

Besides the impact of PFCs on the professionally-exposed populations, recent evidence of pollution from chemical industries producing PFCs have emerged also in the general population from at least four different area worldwide: Mid-Ohio valley in the USA, Dordrecht area in Netherlands, Shandong district in China, and Veneto region in Italy. In the latter, the population at risk includes the cities of Vicenza, Padova and Verona, with 350.000–400.000 potentially exposed subjects, in an area of approximately 150 km² and the consequent contamination of water and food [115]. Despite strong evidence pointing towards a negative role of PFCs on male reproductive function, to date few evidence are available on the actual effect of these substances on seminal parameters in men, with conflicting results [110, 116, 117]. Two cross-sectional studies reported negative associations of PFOS, or high PFOA and PFOS combined, with the proportion of morphologically normal spermatozoa in adult men [116, 118]. Furthermore, in a study of men attending an in vitro fertilization clinic, Raymer et al. [110] reported that luteinizing hormone (LH) and free testosterone significantly and positively correlated with plasma levels of PFOA, although PFOA was not associated with semen quality. Conflicting results are reported also for the association between PFCs and sperm DNA quality, although a significant trend is evident for increased DNA fragmentation in exposed men [117, 119, 120]. In infertile males, PFOS levels were higher than fertile counterparts, together with a higher gene expressions of estrogen receptor (ER) α, ERβ and androgen receptor (AR) [121, 122], suggesting that PFCs activity might be linked also to the genetic expression of sex hormones nuclear receptors. With respect to Androgen Receptor (AR), PFOS and PFOA induce a decrease of the protein expression of this receptor in the hypothalamus and pituitary gland as well as in the testis (Reviewed in [123]). These findings clearly suggest an antiandrogenic potential of PFCs and given the growing evidence suggesting a link between AR disruptors and disorders of male health [124, 125], PFCs effect on AR and consequent derangement of hypothalamic-pituitary axis should be a major concern. However it should be noted that there was a lack of consistent results among the investigated outcomes (Table 1). Subtle associations between higher PFOS and lower testosterone or abnormal semen morphology cannot be excluded. In conclusion, in men, there is little evidence of an association between PFAS exposure and semen quality or levels of reproductive hormones. As is the case for many epidemiological studies, causality cannot be definitively established in these studies, largely because of their cross-sectional design. However the consistency of findings in pre-clinical studies strongly suggests a causal relationship for some endpoints. Some effects are similar to associations seen in humans, whereas other effects cannot be extrapolated to humans given differences in toxicokinetics across species of these chemicals.

Heavy metals have also been recognized as likely inducers of testicular damage and, to this regard, the toxicity of Cadmium (Cd) as environmental contaminant has been known for several decades. Some industrial activities, such as melting and welding of metals, as well as municipal waste incineration are processes that contribute in the release of heavy metals in the environment. Among environmentally exposed population, tobacco smokers are the most exposed subjects, since tobacco leaves accumulate large amounts of Cd, making tobacco smoke the main source of Cd in smokers. Although the mechanisms of testicular toxicity exerted by heavy metals are still under investigation, the permeation through the blood-testis barrier is acknowledged as a fundamental process [126]. Like plasticizers, heavy metals widely employed, in industry as well as in food and dietary supplements [127‐129]. Heavy metals can interfere a the different stages of spermatogenesis resulting in either decrease in sperm count or abnormal increase in sperm counts, sperm DNA damage, and impaired sperm motility [130](Fig. 1). Redox active heavy metals are also found to increase the levels of reactive oxygen species, leading to oxidative stress, induction of DNA damage and apoptosis of spermatozoa together with disruption of the blood-testis barrier and further damaging spermatogenesis [131](Fig. 1). Among heavy metals, Cd has been repeatedly proven to induce reproductive toxicity in the male, which has been extensively reviewed elsewhere [132]. Briefly, experimental studies in animal models strongly support the hypothesis that Cd affects male reproductive function, including spermatogenesis and semen quality, as well as endocrine function (Fig. 1). Indeed, Cd induces severe structural damage to testis vascular endothelium, which ultimately results in necrosis of the testis, and impaired spermatogenesis and testis endocrine function, and affects the BTB integrity, which might lead to susceptibility to toxicity and to the development of autoimmunity against germ cells. Moreover, Cd might induce inflammation and apoptosis within the testis, by means of direct effects on inflammation mediators, and on pro-apoptotic and anti-apoptotic factors, and by interfering with signalling pathways of calcium and cyclic AMP. In addition, Cd exerts targeted effects on selected cell populations of the testis, which include direct cytotoxicity and functional impairment of Sertoli and Leydig cells, and oxidative stress in both somatic and germ cells, mainly by means of mimicry mechanisms and interference with antioxidative activity. Moreover, Cd induces epigenetic modifications in Leydig cells and testis of Cd-treated animals, which might potentially determine an impairmentof semen quality, although these changes were not directly linked to reproductive dysfunction. Lastly, Cd treatment determines a direct disturbance of the hypothalamus-pituitary-gonadal axis (Fig. 1), which might determine the impairment of spermatogenesis and endocrine function (reviewed in [132]). Conversely, evidence from clinical studies is less consistent (Table 1). Observational studies in both environmental and occupational exposed males suggest that Cd have a detrimental effect on semen quality, and can alter endocrine function. Nevertheless, some studies failed to identify differences between Cd exposed and non-exposed subjects, probably due to small-sized study populations, and lack of control for potential confounding variables (reviewed in [132]). Therefore, additional well-designed observational studies, as well as further experimental research in humans, are required to eliminate inconsistencies, and to confirm the effects of Cd on human male reproductive function.