Open access peer-reviewed chapter

Endocrine Disruptors and Reproductive Health in Males

Written By

Tomas Jambor, Hana Greifova, Jana Bistakova and Norbert Lukac

Submitted: February 14th, 2018 Reviewed: May 8th, 2018 Published: October 10th, 2018

DOI: 10.5772/intechopen.78538

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Nowadays, endocrine-disrupting chemicals are considered to be one of the main causes of the ever-increasing occurrence of problems with male fertility. These compounds of natural or anthropogenic origin are omnipresent in the environment and organisms are exposed to them practically nonstop through the air, water, food, and occupationally. Endocrine disruptors have the ability to mimic effects of reproductive hormones and demonstrably can interfere with the endocrine system leading to reproductive disorders at different levels, and considering male reproductive functions, most of the impacts are performed by the breakdown of estrogen- or androgen-mediated processes. A significant body of evidence based upon laboratory or wildlife animal experiments and meta-analysis of semen studies in men indicates that exposure to endocrine disrupting compounds is associated with male reproductive malfunctions, including impairment of spermatogenesis followed by reduced semen quality parameters (sperm concentration, motility, and morphology). Alkylphenols, bisphenol, and phthalates are substantial components of many products with which people come into contact daily. This brief review will emphasize on the possible effects of alkylphenols, bisphenol, and phthalates on the male reproductive system, and current research efforts related to these substances mainly in the context of two main processes taking place in testicular tissues—steroidogenesis and spermatogenesis.


  • male
  • reproduction
  • steroidogenesis
  • spermatogenesis
  • alkylphenols
  • bisphenols
  • phthalates

1. Introduction

Over the last years, many epidemiological studies have been observing worrisome trends in the incidence of human infertility rates. Increasing prevalence of congenital abnorm-alities such as hypospadias and cryptorchidism has also been confirmed by numerous reports. Male fertility generally relies on the quantity and quality of spermatozoa, sufficient activity of Leydig cells, and a proper hormonal balance. Infertility is a widespread problem defined as the inability to conceive after one year of unprotected intercourse. In many cases, there are no obvious signs of infertility. Substantial part of the problem is the disruption of essential cellular processes responsible for normal reproductive functions [1, 2]. Given the short time, genetic changes cannot explain such alterations. We may assume that they only reflect on persistently adverse changes in the environment or in lifestyle. However, it cannot be ignored that some individuals may be more susceptible or resistant to these adverse effects than others, indicating that genetic factors do play key roles [3]. Enormous production and release of industrial chemicals into the environment has led the scientific community to hypothesize that current pollutants may irrefutably disrupt health conditions, leading to extensive damages to physiological functions. In fact, a huge number of chemicals have been found to interact with the endocrine system of different animals in laboratory studies and there is an increasing number of reports on the endocrine disruption in wildlife [4]. Endocrine disruptors (EDs) are an extremely heterogeneous group of ubiquitous synthetic substances, environmental pollutants, and commercial products. They are able to alter functions of the endocrine system, inhibit critical cellular processes, increase the risk of hormone-dependent malignancies, and may result in a wide array of adverse health effects. The term endocrine disruption has been adopted by the vernacular of scientists, toxicologists, and appears here to stay [5, 6]. There are many varied sources of EDs. Typical human exposure occurs with respect to the environmental contamination of the food chain, contact with contaminated household dust, and during the use of personal care products. Other EDs are used as industrial lubricants, solvents and high amounts of EDs were found in household products, pesticides, herbicides, detergents, beverage and food storage containers, metal cans, epoxy resins, etc. Many textiles contain contaminants, such as flame-retardants, including tetrabromobisphenol A and polybrominated diphenyl ethers [7]. Although a chronic exposure to ED takes place through the skin contact or inhalation, the major source are food products. Some experimental studies assume that plastic packaging is the largest source of EDs in the human diet. Repeated exposure of food – contact materials to UV light, acid or alkaline contents and heat may cause polymers to breakdown into monomers as phtha lates, which then leach into the food or beverages [8]. Other by-products such as alkylphenols, bisphenols, polychlorinated biphenyls, dioxins or phthalates are ubiquitous and there is a growing concern that living in an ED-contaminated environment may initiate adverse health effects. Detection of ED residues in human serum, seminal plasma, and follicular fluid has raised concerns that environmental exposure to EDs may be affecting human fertility [9]. Nowadays, some of EDs have been banned or otherwise removed from the industrial processes years ago. On the other hand, these are persistent in the environment throughout many years. A wide range of industrial PCB compounds may be still found in pronounce quantities in the environment, although their manufacture was banned in 1977 [10]. Indeed, humans and wildlife are continually exposed to copious potentially hazardous substances that are released into the environment at an alarming rate.


2. Male reproductive system as a major target of EDs

In this context, possible adverse effects of EDs have been taken into focus, both regarding the effects of EDs on the male reproductive system and with respect to its differential susceptibility towards these compounds. Although there has been an effort to list and rank all possible EDs, the number of evaluated chemicals remains limited. Such information and associated concerns regarding the ubiquitous presences of EDs in the environment have sparked discussions regarding the need for strategies to assess and regulate chemicals with endocrine disrupting properties in order to protect human and wildlife health. During the last years, some epidemiological studies have been comparing an increase in the incidence of male reproductive disorders in many countries. The results showed that the global average sperm count dropped by half and that the sperm motility/viability significantly decreased. In addition, many types of reproductive tract abnormalities were observed in several countries [12]. Several experimental studies have found associations between poor semen quality and increased levels of EDs in the environment [13, 14]. EDs may disrupt not only spermatogenesis, by interfering with germ cells and sperm-supporting cells, but may also affect steroidogenesis occurring in Leydig cells. Many researchers have focused on the potential sources of EDs and their pathological consequences on reproductive health as well as ethnologies in the environment.

2.1. Alkylphenols and their impact on steroidogenesis and spermatogenesis

As we mentioned before, environmental exposure to EDs may adversely affect human and wildlife reproductive functions. Many environmental contaminants including alkylphenols are widely used in the preparation of agrochemicals, industrial and household detergents, paints, and plastics [15]. Alkylphenol ethoxylates, a class of nonionic surfactants, are microbially degraded into alkylphenol diethoxylates and alkylphenol monoethoxylates. These are subsequently degraded into alkylphenols (4-octylphenol; 4-nonylphenol) and along with other subproducts, are known to persist in the environment for a long time [16]. Alkylphenols are endocrine-disrupting agents with native estrogen-like structure and show estrogenic activity. This activity is mediated through the binding of these environmental estrogens to estrogen receptors. Previous studies suggested that estrogenic activity of alkylphenols is linked to a tertiary branched α-carbon and the length of the side chain at that position. Therefore, many experimental studies have investigated estrogen receptor binding and subsequent pathological changes in male reproductive functions. The mechanism also involves interaction with steroidogenic enzymes, transport proteins, and cell signaling processes. However, little is known about the direct effect of alkylphenols on the steroidogenic enzymes (3β-HSD and 17β-HSD) and gene expression [17].

2.1.1. Nonylphenol

One of the most commonly used alkylphenol is nonylphenol (NP). Due to its wide usage, a large amount of nonylphenol is widespread in the environment, especially into water sources. Vazquez-Duhalt et al. [18] have been convinced that the concentration of 0.1 μg/L evokes a public health risk. Based on this knowledge, several studies have investigated the potential impact of NP on male reproductive functions.

Ying et al. [19] demonstrated that nonylphenol’s isomers had different effects on the release of steroid hormones in rat Leydig cells. However, all experimental doses had an unfavorable impact. Specifically, the inhibitory effect of p363-NP isomer was found to be as much as 1.26 times higher than the others. The results imply that the effects of different nonylphenol isomers on the testosterone production do not appear to be completely mediated through the estrogen receptor α or β. For the steroidogenesis, ensured by Leydig cells is an essential conversion of cholesterol into various steroid classes, where 3β-HSD, 17β-HSD, and StAR are responsible for the rate-limiting step. PCR analysis showed that the decrease of testosterone production may be explained by the drastic inhibition of StAR and 3β-HSD gene expression. In a recent study, Wu et al. [20] demonstrated that NP increased testosterone production in rat Leydig cells. The concentration of 127.5 μM NP stimulated the steroidogenic process by elevating the activity of P450scc and stimulating protein expression of StAR. During the same experiment, trypan blue assay was performed. The authors observed the cytotoxic effect of the highest doses of NP (425 μM). Lower experimental doses (42.5–127.5 μM) used in this study had no cytotoxicity until 4 h cultivation. In a previous study, Jambor et al. [21] evaluated the potential impact of NP on the biosynthesis of steroid hormones, cell viability, and ROS production. The production of steroids, specifically dehydroepiandrosterone, androstenedione, and testosterone was reduced following exposure to NP after 44 h of in vitro cultivation. Furthermore, the treatment to NP caused a significant intracellular accumulation of ROS in mice Leydig cells. Majdic et al. [22] reported that NP has an inhibitory effect on P450c17, which is essential in the testosterone synthesis. Several studies demonstrated that NP treatment increased apoptosis of testicular cells, including germ and Sertoli cells [23, 24]. According to Han et al. [25], the highest experimental concentration of NP (250 mg/kg/day) may significantly increase the number of apoptotic cells following in vivo exposure of male rats. Recent evidence also confirms that NP exposure rapidly increases the apoptosis of Sertoli cells in a dose-dependent manner in vitro. The results of flow cytometric analysis indicate that the proportion of apoptotic cells was significantly increased at 20 and 30 μM of NP [26]. Gap junctional intercellular communications (GJIC) were shown to be present between adjacent TM4 Sertoli cells [27]. An important role of GJIC is to regulate cell growth and differentiation and it is also critical for coordinating steroidogenesis and spermatogenesis. Gap junctions are pores composed of connexins (Cx). Several reports indicate that Cx43 is essential for normal testicular functions [28]. Aravindakshan and Cyr [29] showed that the exposure to NP dramatically inhibited GJIC. A significant reduction was observed at 10 μM of NP (almost 80%). The effect of NP on the Cx43 expression was dose- and time-dependent. Time-response analyses in which cells were exposed to 10 μM NP indicated that there was a decrease in Cx43 after 24 h. Exposure of TM4 cell line to NP resulted not only in a decrease in the CX43 levels but also a progressive effect on the level of renewal of the connexins, or on their synthesis, or both was confirmed. In addition, epidemiological studies have reported numerous other adverse effects of nonylphenol on the reproductive system, including reduced testis weights, spermatozoa abnormalities, and a decreased sperm production [30, 31].

NP is considered to be an endocrine disrupting compound which could be involved in declines of both quantity and quality of spermatozoa in adult men [32, 33]. A lot of experiments show an in vitro NP inhibition of sperm motility and viability [34, 35], while in vivo studies confirm spermatotoxicity, spermatogenesis failure, reduced sperm counts and motility, seminiferous tubule degeneration including decreased diameters of seminiferous tubules, lumen and epithelial thickness leading to testicular atrophy [36], and abnormalities in sperm morphology following NP exposure [37, 38]. Huang et al. [39] observed detrimental activity of NP on prepubertal Sprague–Dawley male rats under in vivo and also under in vitro condition, when the animals were treated with 25–100 mg/kg/day for 30 consecutive days by an intraperitoneal injection of NP. NP exposure induced the sperm toxicity, resulting in cell damage and reproductive disorders and initiated oxidative stress, disturbed the PI3K/AKT/mTOR pathway, induced apoptosis and autophagy, and caused developing reproductive damage in vivo and in vitro. Uguz et al. [40] designed an in vitro study with epididymal rat sperm, observed NP-induced (250–500 μg/mL; 1–4 h exposure) impairment of sperm motility, and a decreased mitochondrial membrane potential which probably plays a key role in the malfunction of spermatozoa. Another in vitro experiment with ram and boar spermatozoa provides similar results, when exposure of both sperm types to 250 and 500 μg/mL was harmful to progressive motility, percentages of ram and boar sperm with high mitochondrial membrane potential decreased significantly following exposure to concentrations ≥250 μg/mL. Unlike chromatin integrity, which did not seem to be changed after NP administration, there was a dose-dependent activity of NP on the acrosomal integrity of both species at as low as 1 μg/mL for boar sperm and 10 μg/mL for ram sperm 35]. Lukac et al. [41] used a cell model of bovine spermatozoa to determine the effect of NP (1, 10, 100, and 200 μg/mL) on the motility and viability of spermatozoa during several time periods. The results showed a decreased spermatozoa motility and viability in all experimental samples following the addition of NP after 6 h of exposure. The effects of NP were also evaluated in frozen-thawed bull spermatozoa, when the cells were exposed to concentrations of NP at doses 1, 10, 100, 250, and 500 μg/mL. Sperm parameters were assessed at cultivation times of 0, 1, 2, 3 and 4 h and both motility and mitochondrial membrane potential of sperm cells decreased at concentrations ≥250 μg/mL. In addition, the acrosome reaction was induced even at the lowest concentration of NP [42]. Ergun et al. [43] showed that 100 μg/mL NP induced apoptosis by causing DNA breaks in bovine spermatozoa. Vitellogenesis is a sensitive biomarker of xenoestrogen exposure in vitro and in vivo and vitellogenin is considered to be a key in indicating the presence of xenoestrogens in the environment, as these chemicals have been found to induce the production of this yolk protein in males leading to the impairment of male sexual organ development and disruption of male fertility [44]. NP is estrogenic also to aquatic organisms and experiments related to fish and amphibians have shown that NP is able to induce vitellogenin in the gonads, violating the development of the embryo and larvae, and results in a strikingly skewed sex ratio in aquatic organism via modulating the effects of sex hormones [45]. NP has been connected with the development of different types of sexual dysgenesis in the laboratory and wild fish [46, 47]. Feng et al. [45] investigated the in vivo and in vitro effects of NP on the motility parameters and fertilizing ability of Bufo raddei during amplexus and fertilization period. Based on the results, ROS induced via NP and NP itself was associated with the decrease of the fertilization rate, when in vitro assays showed a direct exposure of sperm to NP with a significant impairment of motility, integrity, and increased ROS levels. Negative correlations were observed between motility of spermatozoa and corresponding ROS concentrations, but the level of NP that admittedly affected spermatozoa in this study (200 μg/L) was about 2.5 times of the highest NP level found in natural aquatic environments (0.065–83 μg/L).

2.1.2. Octylphenol

Numerous reproductive issues such as an increased incidence of testicular cancer, lower spermatozoa activity, and disruption of the steroidogenic process have been related to exposure to alkylphenols. One of the greatly widespread alkylphenols is octylphenol (OP). It is used as a component of emulsifiers, detergents, paints and many other synthetic products. Nowadays, OP is mainly present in sediments, surface waters, and even drinking water. Due to its relative stability and hydrophobic properties, OP is bioaccumulated in various tissues and poses a large health risk for the organism [48, 49, 50]. It has been reported that certain doses of OP may negatively affect cellular processes such as steroidogenesis and spermatogenesis essential for a normal development and functions of the male sex. However, there are still limited information about the mechanism, through which OP affects biosynthesis of steroid hormones. Some experimental studies have hypothesized that OP may directly modulate the activity of steroidogenic enzymes. Murono et al. [51] documented that exposure to 2000 nM OP affected the testosterone production in rat Leydig cells. In response to the experimental dose, testosterone levels significantly increased after 2, 4, and 8 h cultivation, when compared with the control. Exposure to shorter periods (0.5 and 1 h) were also examined; however, the weak increase at these times was not statistically significant. The increase in hormone production was not associated with changes in cAMP levels and it did not involve the estrogenic activity (binding) to the estrogen receptors. Furthermore, higher testosterone secretion was not the consequence of inhibiting 5α-reductase activity in Leydig cells. Although these results did not describe signaling pathways, it is necessary to identify the potential mechanisms through which intermediate stages of steroidogenesis may be affected. Some epidemiological studies imply that the inhibiting effects of OP on the steroidogenesis are mediated through the potential of OP to generate ROS and inhibit testosterone secretion. Cytochrome P450scc and P450c17 are essential in converting cholesterol to testosterone in Leydig cells. During the steroidogenic process, ROS are produced by electron leakage outside the electron transfer chains and these radicals may cause lipid peroxidation to inactivate P450 enzymes [52]. Several reports evaluated the potential effects of OP on the steroid hormone synthesis [51, 53]. According to Kotula-Balak et al. [54], independently of the incubation time, high doses of OP significantly inhibited the progesterone production in mice MA-10 cells. Inhibition in progesterone levels was significantly higher in the experimental groups cultivated with OP for 3 h than in cells incubated for 12 h. This can be related to the restoration of Leydig cell steroidogenic function within the time of culture. Decreased progesterone production could be mediated through the inhibition of 3β-HSD since it was reported that estradiol inhibits the progesterone level via the disruption of the 3β-HSD function. Murono et al. [55] investigated the impact of OP on the biosynthesis of steroid hormones in rat Leydig cells in vitro. The authors reported a biphasic effect, where the lower experimental doses (1 and 10 nM) increased the testosterone production by approximately 10–70% above the control group, whereas higher concentrations (100 and 200 nM) decreased the testosterone level progressively. The inhibitory effect of OP was also evaluated by Nikula et al. [53]. Inhibition of testosterone secretion by 4-t-octylphenol in cultured mice Leydig cells has been suggested to occur at the 17β-HSD step. It has also been reported that the gestational exposure of pregnant rats to OP decreases the amount and activity of the P450c17 steroidogenic enzyme in male offspring and SF-1 (steroidogenic factor) involved in the gonad development and expression of steroidogenic enzymes [56]. Based on the evidence gathered from the literature, it seems possible that inhibited functions of a male reproductive system might be mediated not only through the disruption of steroidogenic enzymes but also via the direct toxic effect of OP, resulting in a lower cell viability and apoptosis. Qian et al. [50] evaluated the cytotoxic effect of OP (30–60 μM) in rat Sertoli cells after a 24 h exposure. Cell viability was significantly reduced at 40, 50, and 60 μM OP. Additionally, the highest experimental dose decreased the Sertoli cell viability in a time-dependent manner with a significant decrease following a 12 h cultivation. The cytotoxic effect of OP is strongly dependent on the experimental doses. Jambor et al. [57] evaluated the in vitro effect of 4-OP on mice Leydig cell viability. The results showed a greater viability at 1, 2.5 and 5 μg/mL of 4-OP following 44 h of cultivation. Kotula-Balak et al. [54] illustrated marked differences in the Leydig cell morphology after OP treatment. Mice Leydig cells exposed to experimental doses of OP (10−4 to 10−6 M) grew in a small group and 60% of cells showed nucleus shrinkage, cytoplasm vacuolization and membrane floating, while the control cells were formed as a monolayer with an epithelioid shape and abundant cytoplasm. Conversely, lower concentrations of OP did not markedly affect the morphological structure of exposed cells. In the recent years, a link was confirmed linking OP and the increased incidence of male reproductive dysfunction. The ability of OP to affect spermatogenesis has been the subject of much investigation. Spermatozoa abnormalities, a decreased sperm motility and lower spermatozoa viability are current problems mediated through OP exposition. Of the alkylphenols examined for their ability to act as an estrogen compound, octylphenol has been observed to be vastly effective, showing approx. one thousandth of the estrogenicity when compared to a strong estrogen 17β-estradiol [58]. Exposure to OP extremely inhibits the testicular function as exhibited by a reduced size of the testes, reduced androgen concentrations, and a negatively affected spermatogenesis. Similarities in the activity of OP and those noticed after the addition of 17β-estradiol indicate that OP exerted its effect to impair the testes in an estrogenic-like manner on the hypothalamus and/or anterior pituitary gland to arrest the gonadotropin secretion [59]. OP is also believed to support the reduction in sperm quantity in men resulting in male infertility and it has been defined as a potential reason of reproductive tumorigenesis [60]. It has also been reported that OP shows a toxic potency on cultured prespermatogonia and Sertoli cells [61]. In addition, it is proved that OP is able to generate ROS which are cytotoxic compounds resulting in oxidative damage associated with damage to biomolecules such as membrane lipids and DNA in sperm cells [62]. Adverse effects of OP on male reproductive functions in pubertal rats were evaluated by Herath et al. [63], when 50-day-old rats in the OP group received daily injections of the xenoestrogen at a concentration of 3 mg/kg. After 5 weeks of exposure, the epididymal sperm motility and sperm head counts were determined with reduced sperm counts resulting from a decreased plasma testosterone, but without significant effects of OP on the sperm motility parameters. The potential in vivo genotoxic activity of OP in adult male Wistar albino rats was studied by Ulutas et al. [64], when animals received OP oral doses of 125 and 250 mg/kg for 4 weeks. Possible genotoxic effects of OP were evaluated as comet parameters including tail length and tail moment with significant differences in both tested parameters only in the case of animals treated with the highest dose of OP. Peng et al. [65] also provide results of a combined genetic toxicity of OP along with NP in male mice following a peritoneal injection of nonylphenol-octylphenol (50, 100, and 200 mg/kg). The effect on the DNA damage in the testicular cells and sperm deformation rate after the exposure were measured using the comet assay and sperm morphologic test. Within the examined doses of 100 and 200 mg/kg, the quantity of the comet cells in the testes cells was increased. The DNA migration length was also significantly increased as OP-NP elevated and the rate of sperm deformation was higher following exposure to the tested chemicals too [66]. OP was also examined in the context of the biochemical composition of the seminal fluid and production of the viviparous eelpout (Zoarces viviparus) and the investigation was carried out at the time of spawning. After 10 days of exposure to OP, a decline in the gonadosomatic index was observed following the milt volume with a spermatocrit increase. The histological investigation manifested that OP impaired the lobular composition, including the Sertoli cells. In most of the OP-exposed individuals, trapped sperm cells in parts of the seminiferous lobules and the sperm ducts were observed. OP also affected the biochemical composition of the seminal fluid with elevated concentrations of the tested parameters such as magnesium, calcium, and total protein, meanwhile values of free amino acids were decreased in the exposed fish [67]. Movement characteristics are always the most important parameters in the evaluation of semen quality. Spermatozoa motility represents the primary characteristic in the assessment of male fertility and it is a fundamental premise for a successful fertilization. Motility parameters are closely linked to the mitochondrial activity of spermatozoa as these organelles play a key role in the energy provision by production of ATP [68]. Lukacova et al. [69] confirmed a decline of bovine sperm motility, progressive motility, and mitochondrial activity after exposure to 1–200 μg/mL OP during several time periods (0, 2, 4, and 6 h). Interestingly, the values of intracellular superoxide production revealed a slight decline of the superoxide concentration at the dose of 1 μg/mL when compared to the control group and conversely, doses 10, 100, and 200 μg/mL of OP increased the concentrations of superoxide in bovine sperm. Thus, in general, the effects of alkylphenols on the testicular function are not clearly defined yet and their effect may be attributed to the concentration, estrogen-mimicking activity, and time of exposure.

2.2. Bisphenols and their impact on steroidogenesis and spermatogenesis

Exposure to xenoestrogens such as bisphenols has been shown to cause adverse effects on male reproductive system in humans and numerous animal species. As typical endocrine disruptors, bisphenols are one of the most studied xenoestrogens in the field of male reproductive system. A survey of the Pubmed database provides more than 10,000 articles on the topic, including epidemiological as well as experimental studies. The overwhelming majority of bisphenols is used as stable components of household products, epoxy resins, inner surface of food metallic cans, dental sealants, and for myriad additional synthetic products. Many of us are mostly confronted by bisphenols through gastrointestinal exposure (food packaging) and dermal exposure (paper money and paper products). It is well known that increased concentration of bisphenols was detected in urine, milk or sweat and over 90% of human population is daily exposit to bisphenol A. Subsequent bioaccumulation and kinetic properties may adversely affect the overall health [70, 71]. Nowadays, bisphenols have been associated with a variety of human diseases, specifically kidney and cardiovascular diseases, obesity, developmental defects, and reproductive disorders. Recent studies indicate a direct link between the incidence of male reproductive dysfunction and rising concentrations of bisphenols in the environment. A decrease in semen quality was the first reported alteration and from this moment on an informative expansion was launched on the potential consequences of bisphenol exposure [72]. Several reports demonstrate a direct effect of bisphenols on the biosynthesis of steroid hormones. Negative effects of bisphenol A (BPA) have been reported in both in vivo and in vitro studies, where the steroidogenic enzymes were recognized as primary targets. Downregulation of the expression levels of CYP11A and CYP17A has been observed primarily, resulting in the decline of testosterone synthesis [73]. The altered levels of testosterone may cause subsequent reproductive dysfunction by interfering with the feedback regulatory mechanisms. Another serious effect by which bisphenolic compounds perform their adverse impact on the male reproductive cells are disruption of the brittle balance between the antioxidant capacity of cells and prooxidants in testicular tissues, which is linked to the increased risk of oxidative stress development resulting in the arrest of spermatogenic processes, production of abnormal sperm cells, and impairment of normal existing sperm cells in the reproductive tract [11]. Oxidative reactions may lead to the decline of spermatozoa quality, as observed by the decrease of spermatozoa motility, velocity, and viability values. Moreover, bisphenol exposure could also result in the depletion of ATP metabolism and damage to the genetic material by sperm DNA fragmentation [74].

2.2.1. Bisphenol A

Lan et al. [75] evaluated the effects of BPA on two steroidogenic enzymes (CP11A1; CYP19) essential for the normal steroidogenic process. According to the PCR analysis, the endogenous gene expression in both was upregulated by BPA at 100–1000 nM. Another steroidogenic enzyme, CYP17, involved in the testosterone synthesis was also measured. The results showed that BPA did not affect CYP17 protein expression significantly. However, the authors hypothesized that the balance of steroid hormones may be affected. This was confirmed in the next part of the study, where the testosterone production was slightly decreased at 1–100 nM BPA following a 24 h exposure. The next steroidogenic enzyme responsible for the conversion of pregnenolone to progesterone is 3β-HSD. Ye et al. [76] reported a significant inhibition of the 3β-HSD activity in rats and humans. Human 3β-HSD was more sensitive to BPA’s inhibition than the rat enzyme. The authors also evaluated the effects of BPA on the testosterone production in rat Leydig cells. Experimental doses of 10 and 100 μM markedly decreased the testosterone generation. Importantly, evidence exists that exposure to BPA in utero may reduce the neonatal serum testosterone level [77]. In summary, although BPA directly affects the steroidogenic genes, it is clear that BPA disrupts the hormone synthesis and contributes to reproductive disorders. Because of an increased concern over the safety of BPA, European Union has banned its use in plastic bottles for infants. The viability of Leydig cells is a significant indicator for a sufficient production of steroid hormones. Lan et al. [75] illustrated a dose-dependent effect of BPA on this parameter in the MA-10 cell line. The data show a decrease in the cell viability (1–200 μM) following a 24 h cultivation in vitro. However, significant differences were recorded only with respect to the highest dose of BPA (200 μM). Goncalves’s et al. [78] study showed a decrease in the Leydig cell viability upon the exposure to BPA. The authors found out that experimental doses above 1 μM inhibited the cell viability following a 24 h incubation compared to the control. Nonetheless, the viability of TM3 cell line did not decrease significantly even after a 48 h exposure at concentrations below 50 μM. De Freitas et al. [79] observed a significant reduction in the viability of human Sertoli cells after the cultivation with 10 μM BPA for 48 h.

Nowadays, there are many epidemiological studies which evaluated the effect of bisphenols on the spermatozoa or spermatogenesis. Observable changes were recorded in the spermatozoa motility, spermatozoa viability, and DNA integrity. In vivo experiments with adult male rats indicated that the low concentration of BPA (2 μg/kg body weight) administered orally can effectively inhibit spermatogenesis via disruption of the biosynthesis of reproductive hormones resulting in the meiosis inhibition of sperm cells and induction of the Fas/FasL pathway with a subsequent apoptosis. Declining amounts of testosterone were followed by a reduction of sperm quantity [80]. Evidence showed an obvious link between increased urine levels of BPA and reduced values of the sperm concentration what can be attributed to the disturbed processes of spermatogenesis following BPA exposure. Harmful effects of BPA on the spermatogenesis observed in experimental animals are also in agreement with an epidemiological study focused on the impact of BPA on exposed human males. Reduced spermatozoa count, indicating a primary association between BPA exposure and production of sperm cells were attributed to increased values of BPA in urine when men with high urine BPA levels had more than three times lower sperm concentration and viability; however, no correlation was observed between the urine BPA concentrations and semen volume or abnormal sperm morphology compared to subjects without the presence of BPA in the urine [60]. Also, other in vitro studies revealed a direct effect of BPA exposure on the sperm quality. Singh et al. [81] used in his in vitro study chicken sperm to determine environmentally relevant concentrations of BPA (0.18, 0.37 and 0.74 mM) related to motility, fertilizing ability, live sperm percentage, and mitochondrial membrane potential after 30 min of BPA treatment. The results showed that 0.74 mM BPA is able to compromise sperm functions in the case of all analyzed parameters leading to the decline of sperm fertilizing ability. Data obtained from in vitro experiments by Lukacova et al. [82] refer that BPA has negative effects on bovine spermatozoa motility in different doses (1, 10, 100, and 200 μg/mL). The results showed that BPA has the ability to reduce the values of mitochondrial activity and spermatozoa motility, causing mitochondrial damage as evidenced by the increased values of intracellular superoxide. Spermatozoa motility parameters were significantly decreased in experimental groups exposed to concentrations of BPA higher than 100 μg/mL. In experimental mice, the motility analyzed following 6 h of in vitro treatment with 0.0001, 0.01, 1, and 100 μM BPA, the number of motile sperm cells was also reduced in the case of dose of 100 μM BPA [83]. Administration of different BPA concentrations (0.6, 4.5, and 11.0 μg/L) demonstrated an impairment of motility in fish spermatozoa too [84].

2.2.2. Bisphenol alternatives

More stringent global regulations of BPA production and the use have led to the development of alternative bisphenol compounds [85]. A few years ago, researchers have begun to deal with potential properties of 4,4′-dihydroxydiphenylsulphone (BPS) or 4,4′-dihydroxydiphenlymethane (BPF). Both are presently not regulated and are used without restriction. Additionally, currently available toxicological data are scarce and the information about their potential impact is limited. Nowadays, studies reported the effects of BPS via genomic mechanisms using extremely high concentrations but there are still no studies evaluating the in vivo toxicity. Although BPS is less likely to leach from plastic packaging with heat, it does still escape the polymer in small quantities under the normal use. Chen et al. [86] showed that 40 μM BPS had a 15-fold lower genomic estrogenic activity than BPA. Only a few studies have evaluated BPS at low concentration ranges likely to be present in foods, wildlife or humans. Eladak et al. [87] used the mouse FeTA model to illustrate the effects of BPS and BPF on the testosterone synthesis. Results from the present study showed that BPF has a similar dose-response effect as BPA with a significantly decreased amount of testosterone starting from 1000 nmol/L. On the other hand, BPS had even a more potent inhibitory effect than BPA. Indeed, 100 nmol/L BPS significantly reduced the testosterone production after 3 days of treatment. Authors also compared the effects of 10,000 nmol/L BPA, BPS, and BPF on specific gene expression in mice Leydig cells. All bisphenol alternatives reduced the expression of key genes involved in steroidogenesis such as Star, hsd3βa CYP17a1, expect CYP11a1. In addition, the expression of Lhcgr (the gene encoding the LH/CG receptor) was also decreased. This is one of the few reports that suggest harmful consequences on the reproductive functions in humans and rodents. According to Ji et al. [88], BPS is able to reduce the level of testosterone as well as CYP17a and 17β-HSD mRNA levels in zebrafish. It must be noted that the binding activity of BPS and BPF to estrogen receptor (α; β) is, respectively, 5-or 10-fold lower than that of BPA in the HELN cells [89].

Effect of BPS exposure on oxidative stress, generation of ROS, and impairment of DNA integrity of rat sperm cells under the in vitro condition and daily sperm production and sperm DNA damage under the in vivo condition was examined in the study of Ullah et al. [90]. Spermatozoa were cultivated along with BPS at doses of 0.5, 1, 10 and 100 μg/L and the analyses showed that the highest concentration of BPS initiated ROS generation, induced peroxidation of membrane lipids, altered superoxide dismutase concentrations, and increased the incidence of DNA fragmentation in the sperm cells. The in vivo part of this study revealed that adult rats exposed to concentrations of 0.5, 5, 25, and 50 μg/kg/day for 28 days demonstrated a decline in daily sperm production with rising values of DNA damage occurring in spermatozoa observed in experimental animals treated with the highest dose (50 μg/kg/day) of BPS; however, the motility parameters were not inhibited. Similarly, treatment with 50 μg/kg/d lead to the development of oxidative stress in the testes and impaired reproductive functions in rats [91]. An earlier study on zebrafish embryos focusing on the developmental exposure to BPS was performed to examine the reproduction potential and hormonal balance in adult individuals. Embryos of zebrafish were treated and bred in the presence of various doses of BPS (0, 0.1, 1, 10, and 100 g/L) for 75 days. Following that period, adult males and females were paired for next 7 days in fresh water and subsequently the impact on individual development, reproduction, plasma vitellogenin, sex steroids, and thyroid hormone rates were examined. The results showed skewed sex ratio in favor of females and decreased values of body length and weight in males exposed to 100 g/L of BPS. The gonadosomatic index showed reduced values in fish at tested concentrations ≥10 g/L of BPS. In both males and females, a significant stimulation in plasma vitellogenin level was noticed at ≥10 μg/L of BPS and also thyroxine and triiodothyronine levels were significantly decreased at 10 and 100 μg/L of BPS in males. Sperm count was also reduced in the experimental groups exposed to 10 and 100 μg/L of BPS [92]. In other studies, cytotoxic, genotoxic [93], and mutagenic [94] effects of BPS in different cell models were documented. It is proved that the exposure to BPS can violate the cellular signaling path in the apoptotic and viability ways, which is why it is possible to expect a reaction of BPS with pro-apoptotic and signaling cascades observed also in the sex cells resulting in the affected cell cycle and apoptosis [95]. Nowadays, further research is required to elucidate the effects of bisphenols on the male and female reproductive system.

2.3. Phthalates and their impact on steroidogenesis and spermatogenesis

Numerous environmental contaminants have hormonal or anti-hormonal actions that interfere with endocrine homeostasis of individuals. As we mentioned above, the group of endocrine disruptors is very heterogeneous and phthalates, as ubiquitous chemical compounds are widely used as plasticizers in children’s plastics toys, food packaging, medical tubing, certain cosmetics, shampoos, soaps, and many others household products [96]. Early experimental studies found a low level of phthalate toxicity in rodents, but nowadays, a high extent of carcinogenicity, teratogenicity or testicular atrophy has been widely confirmed. Recent studies have verified that phthalates are capable to affect many physiological mechanism and functions, especially within the reproductive system. Moreover, disorders linked to reproductive toxicity may appear in early life stages, puberty, and some of them may manifest in adulthood. The Department of Health and Human Services estimated that daily human consumption of commonly used phthalates diethylhexyl phthalate (DEHP) revolves around 5.8 mg and monoethylhexyl phthalate (MEHP) ranges from 3.26 to 4.15 in males and 2.93 to 3.51 in females. On the other hand, DEHP is metabolized by intestinal lipases to MEHP, which is glucuronized and excreted from the organism with minimum tissue accumulation [97, 98]. According to its toxicological profile, MEHP seems to be 10-fold more potent in its toxicity to Leydig and Sertoli cells in comparison to DEHP, suggesting that DEHP is the pretoxin which acts via metabolizing into MEHP [99]. Several toxicological reports suggest that DEHP and MEHP disrupt reproductive development and now it is established that these phthalates inhibit the biosynthesis of steroid hormones in Leydig cells at different developmental stages. In utero exposure to phthalates has been shown to reduce male fertility potential in rats. Subsequent postnatal changes preceded an inhibition in Leydig cell function, including lower levels of testosterone. Many authors suggest that phthalates exert their effect via multiple mechanism of action such as the peroxisomes proliferator-activated receptors, estrogen receptors or yet unidentified mechanism.

2.3.1. Diethylhexyl phthalate (DEHP)

Akingbemi et al. [100] investigated the ability of DEHP to affect the biosynthesis of steroid hormones in rat Leydig cells. Pubertal rats were exposed to 1, 10, 100, and 200 mg/kg/day DEHP for 2 weeks. The highest experimental dose (200 mg/kg/day) DEHP caused a 77% decrease in the activity of 17β-HSD and reduced the testosterone production to 50% of the control. Paradoxically, prolonged time of cultivation to 28 days resulted in significant increases in the testosterone secretion capacity and in serum LH levels. A few years later, Akingbemi et al. [101] evaluated the potential effects of DEHP on isolated rat Leydig cells in vitro. When compared to the control, mRNA levels of PCNA and cyclin D3 were expressed at statistically higher levels of proliferation following treatment. Additionally, estradiol levels were elevated by as much as 50% above the control group and aromatase gene expression was also higher in DEHP exposed cells. Several recent investigations have shown that DEHP disrupts the reproductive system of the male rat in an antiandrogenic manner. In the present study, Parks et al. [102] explored the antiandrogenic action of DEHP and MEHP as well as alterations in the testosterone production. Maternal exposure at 750 mg/kg/day caused a significant reduction in the testosterone levels. In addition, Liu et al. [103] performed gene expression profiling following in utero exposure to phthalates and observed a decline in levels of steroidogenic enzymes (CYP11a1; CYP17a1) and lipid transport (StAR). However, the exact mechanism of action is not fully clear. The negative impact of DEHP on the male reproductive system has been related to their monoester metabolite MEHP. It has been shown that this endocrinologically active phthalate may negatively affect the testes and more specifically suppress Leydig cells functions [104].

2.3.2. Monoethylhexyl phthalate (MEHP)

Dees et al. [105] reported that MEHP inhibits androgen production in MA-10 Leydig cells. By using different MEHP concentrations over a longer time interval (24 and 2 h), the authors have demonstrated that even at low experimental doses MEHP inhibits the steroid production (a 50% inhibition was observed at 10 μM), induces morphological changes such as mitochondrial swelling and vesiculation of the Golgi apparatus. Conversely, at 100 and 300 μM doses, this inhibition was not seen. Thus, it is possible that the absence of any effect may be mediated through an unidentified mechanism, distinct to the mechanisms responsible for the inhibition of steroid production. In the next in vitro study, Jones et al. [106] exposed the primary culture of Leydig cells to MEHP (1 mM) for 2 h. A moderate decrease in testosterone production was shown which correlated with the changes in the cell ultrastructure. Treatment with MEHP confirmed mitochondrial swelling with the loss of matrix granules, reduction in the number of Golgi apparatus and dilatation of the smooth endoplasmic reticulum. Svechnikov et al. [107] also confirmed the inhibitory effect on steroidogenesis in rat Leydig cells. The result showed significantly lower testosterone levels (57–62% inhibition) in exposed cells (250 μM MEHP) after 24 h incubation when compared with the control group. In order to determine whether the inhibition of testosterone secretion was due to the disruption of StAR, the authors decided to monitor the expression of this protein by Western blotting. A marked decrease in StAR expression was observed after 24 h incubation. In addition, the activity of 5α-reductase, an enzyme synthesizing the potent androgen dihydrotestosterone, was dramatically inhibited in immature Leydig cells. The dysfunction of Leydig cells is postulated to have a direct association with androgen-dependent parameters of sexual development. Nevertheless, it is necessary to determine whether the effects of chronic DEPH or MEHP exposure are reversed or mitigated when exposure is terminated.

Numerous studies have evaluated the testicular toxicity of phthalates in different experimental models and showed that spermatozoa and spermatogenesis were one of the main targets of their actions. Kasahara et al. [96] indicate associations between DEHP administration and increased production of ROS and selectively decreased GSH and ascorbic acid in the testis with a consequent induction of rat sperm cell apoptosis leading to testicular atrophy after in vivo DEHP exposition. More specifically, the results provided by Li et al. [108] when male rats were fed DEHP for 2 weeks. The result was that the spermatogenesis became disrupted with decreased spermatocytes and spermatids counts and in addition, DEHP (20, 100, 500, to 1000 mg/kg) appeared to inhibit DNA replication. This resulted in the induction of the mitochondrial apoptotic pathways and overgeneration of ROS. Also, elevated activity of superoxide dismutase, reduced activity of glutathione peroxidase, and increased values of malondialdehyde after exposure to 500 mg/kg/day of dibutyl phthalate in the epididymis support the importance of oxidative stress as a major mechanism of phthalate action [109]. Likewise, the response to oxidative stress indicates an increased expression of mitochondrial peroxiredoxin and cyclooxygenase-2 in germ cells after phthalate treatment [110]. Apoptosis of germ cells has also been proposed as a potential effect of phthalates on male reproduction based on the results reporting an increased membrane localization of Fas and apoptic cells [111, 112]. One essential trace element necessary in spermatogenesis is zinc and even a slight deficiency of zinc has been observed to arrest spermatogenesis in both mice and humans [113]. Earlier studies examined phthalate-induced modifications in metabolism of zinc after treatment with high doses of phthalates with reduced testicular zinc concentrations [114], a decline of zinc half-life in the testes [115] and increased excretion of zinc in urine after phthalate exposure [114, 116]. The next schematic figure (Figure 1) summarizes final findings.

Figure 1.

The effects of selected EDs on the male reproductive system.


3. Future directions and recommendations

Probably, research is just at the very start of a long journey to refine understanding of the principal mechanisms of toxicity related to endocrine disruptive compounds and the range of influence of these hormonally active substances to the human and environmental health in the context of male reproduction. Society will definitely continue to use these materials because of their undeniable benefits and primary we have to aim future investigation on testing and development of chemicals to maintain healthier, safe, and more sustainable world for next generations and on evolve suitable strategies of remediation of EDs. Progress in the experimental area of endocrine disruptors effects provides rich lessons that can be usable in other fields of science, as well as in the future missions in toxicology and environmental health.

This still controversial and live topic has already improved research of toxicology and risk assessment and has moved it into certain radically different trends. Further improvement in this field including reproductive biology rests in modern technology, such as toxicogenomics, which can study precursor changes on the level of cells and biological molecules and thus offer understanding of dose and time-dependent responses in more detail. Moreover, the increased usage of human, rather than animal, cell models keep a promise for intensify issues of human relevance. However, reality is that new questions are asked while previous issues associated with impact of EDs on male reproductive organs and behavior persist. The most important fields of investigation for better understanding of how EDs affect functions of tissues involved in male reproductive physiology are associated especially with questions such as why are some tissues, time periods, and even organisms more resistant to EDs exposure; how EDs effect in model organisms and cells translates to human exposure to EDs. There is also need for more studies with aim on syndromes and EDs contribution to development of multiple symptoms at once. The summary of some EDs affecting male reproductive system is presented in Table 1. There is also necessity to interpret specific cell culture responses in the context of whole-organism physiology, ideally that of humans. It is well known that endocrine system mediates reactions on distant tissues and cells. Therefore, research that focuses only on isolated components of endocrine system or target tissues may provide incomplete information. Essential principles of toxicokinetics should be part of key studies related to impact of EDs on specific structures of organisms.

Chemicals Cellular effects Source/applications Study
Aldrin Competitive binding to androgen receptors; ↓weight of testes; ↓ 3β-HSD and 17β-HSD; ↓spermatozoa MOT; Insecticide, groundwater Lemaire et al. [117]
Chatterjee et al. [118]
Das Neves et al. [119]
Alachlor Competitive binding to estrogen and progesterone receptors; no effects on testosterone production;
↓ spermatozoa MOT and viability;
Herbicide Mikamo et al. [120]
Gizard et al. [121]
Bisphenols Estrogenic and anti-androgenic affinity; ↓ 3β-HSD and 17β-HSD; ↑apoptosis; ↓ sperm MOT, viability and concentration; Plasticizers, epoxy resins, dental sealants, Eladak et al. [87]
Lukacova et al. [82]
Akingbemi et al. [122]
Ahmed [123]
DDT and metabolites Competitive binding to androgen receptors, activation of androgen-sensitive cells proliferation;
↓ expression of steroidogenic enzymes;
↓ testosterone, estradiol, progesterone production;
Pesticides, insecticide Tapiero et al. [124]
Tesier and Matsumura [125]
Castellanos et al. [126]
Mono/Di-(2-ethylhexyl) phthalate ↓17β-HSD; ↓ StAR expression,
↑ mitochondrial damages; ↑ ROS;
↓ antioxidant defense; ↑spermatozoa apoptosis;
Plasticizers, cosmetics, food packaging Akingbemi et al. [100, 101]
Svechnikov et al. [107]
Dees et al. [105]
Alkylphenols ↓3β-HSD, 17β-HSD, StAR; ↑ ROS production; ↓ cell viability;
↑ apoptosis; ↓ spermatozoa MOT and viability; ↑ DNA fragmentation,
Cosmetics, pesticides, paints, food packaging’s, Jambor et al. [21, 57]
Lukacova et al. [69, 82]
Diemer et al. [127]
Haavisto et al. [128]

Table 1.

Summary of some EDs affecting male reproduction.

3β-HSD, 3beta-hydroxysteroid dehydrogenase; 17β-HSD, 17beta-hydroxysteroid dehydrogenase; MOT, motility; StAR, steroidogenic acute regulatory protein; and ROS, reactive oxygen species.


4. Conclusion

In recent years, a growing incidence of EDs has led scientific community to show how these substances may affect the male reproductive system. The in vitro evaluation of steroidogenesis and spermatogenesis are necessary for the screening potential of reproductive toxicants such as alkylphenols, bisphenols, phthalates, and many others. The mechanism of their negative effect is by diverse but one important endpoint is reduced processes, essential for normal reproductive functions. This review has demonstrated that certain groups of EDs may directly or indirectly interfere with the biosynthesis of steroid hormones and spermatogenesis via different mechanisms of action. Dysfunction of these processes may cause an incomplete masculinization, suppressed libido, reduced steroidogenic capacity, develop various malformations in spermatozoa and subsequently totally inhibit the reproductive potential of humans and animals. It must be noted that further studies are required to understand the effects of EDs on the male reproductive functions and their contributions to male sub- or infertility.



The study was supported by the Slovak Research and Development Agency Grant no. APVV-16-0289, APVV-15-0543, APVV-15-0544, VEGA 1/0539/18, and KEGA 010SPU-4/2018.


Conflict of interest

The authors declare no conflicts of interest.



ARandrogen receptor
AhRaryl hydrocarbon receptor
cAMPcyclic adenosine monophosphate
ERestrogen receptor
PCBpolychlorinated biphenyl
PCRpolymerase chain reaction
ROSreactive oxygen species


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Written By

Tomas Jambor, Hana Greifova, Jana Bistakova and Norbert Lukac

Submitted: February 14th, 2018 Reviewed: May 8th, 2018 Published: October 10th, 2018