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Arowolo, O.;  Pilsner, J.R.;  Sergeyev, O.;  Suvorov, A. Male Reproductive Toxicity of Polybrominated Diphenyl Ethers. Encyclopedia. Available online: (accessed on 19 April 2024).
Arowolo O,  Pilsner JR,  Sergeyev O,  Suvorov A. Male Reproductive Toxicity of Polybrominated Diphenyl Ethers. Encyclopedia. Available at: Accessed April 19, 2024.
Arowolo, Olatunbosun, J. Richard Pilsner, Oleg Sergeyev, Alexander Suvorov. "Male Reproductive Toxicity of Polybrominated Diphenyl Ethers" Encyclopedia, (accessed April 19, 2024).
Arowolo, O.,  Pilsner, J.R.,  Sergeyev, O., & Suvorov, A. (2022, November 23). Male Reproductive Toxicity of Polybrominated Diphenyl Ethers. In Encyclopedia.
Arowolo, Olatunbosun, et al. "Male Reproductive Toxicity of Polybrominated Diphenyl Ethers." Encyclopedia. Web. 23 November, 2022.
Male Reproductive Toxicity of Polybrominated Diphenyl Ethers

Polybrominated diphenyl ethers (PBDE) are a group of flame retardants used in a variety of artificial materials. Despite being phased out in most industrial countries, they remain in the environment and human tissues due to their persistence, lipophilicity, and bioaccumulation. Populational and experimental studies demonstrate the male reproductive toxicity of PBDEs including increased incidence of genital malformations (hypospadias and cryptorchidism), altered weight of testes and other reproductive tissues, altered testes histology and transcriptome, decreased sperm production and sperm quality, altered epigenetic regulation of developmental genes in spermatozoa, and altered secretion of reproductive hormones. A broad range of mechanistic hypotheses of PBDE reproductive toxicity has been suggested. Among these hypotheses, oxidative stress, the disruption of estrogenic signaling, and mitochondria disruption are affected by PBDE concentrations much higher than concentrations found in human tissues, making them unlikely links between exposures and adverse reproductive outcomes in the general population. Robust evidence suggests that at environmentally relevant doses, PBDEs and their metabolites may affect male reproductive health via mechanisms including AR antagonism and the disruption of a complex network of metabolic signaling.

PBDE dose–response relationships mTOR signaling steroidogenesis metabolic disruption sperm epigenome

1. Introduction

Polybrominated diphenyl ethers (PBDEs) are a group of halogenated biphenyl chemicals that structurally contain bromine atoms [1]. Depending on the number and location of the bromine atom(s) on aromatic rings, PBDEs are classified into 209 congeners [2]. This group of chemicals has been used extensively as flame retardants in various consumer products such as construction, textile, electrical, plastics, and furniture materials [3][4]. Globally, three major commercial mixtures of PBDEs have been produced and used: deca-, octa-, and penta-BDE [5]. Historically, the use of PBDE in various forms dates to the 1960s; it was used continuously for various purposes for almost five decades before most congeners were banned [6]. During the Stockholm Convention in 2009 and 2017, deca-BDE, octa-BDE, and penta-BDE were added to the list of persistent organic pollutants (POPs) that should be eliminated from the environment [7]. The highest demand and production of total PBDEs was documented to be in the year 2003 and reached approximately 90 kt per year [8]. Penta- and octa-BDE were voluntarily removed by industry from the USA market by 2004 [9], while deca-BDE was removed by 2013 [10]. Studies have shown that levels of PBDEs in the United States have been and remain higher than in any other country in the world [11].
PBDEs escape to the environment during their production, utilization, recycling, and disposal; this is due to their characteristic nature of not binding to complex polymers [12]. PBDEs are highly ubiquitous, persistent in nature, and can travel long distances [4][13][14]. The high lipophilicity of PBDEs (Log Kow of 5.74–8.27) is responsible for their persistence, bioaccumulation, and biomagnification [15]. Humans are usually exposed to PBDEs via the inhalation and ingestion of dust containing PBDEs, and the ingestion of contaminated food [16]. The half-lives of PBDE congeners in the human body are estimated to be between 1 and 12 years [17][18]. The highest levels of PBDEs were detected in human adipose tissue samples collected in New York from 2003–2004 and ranged from 17 to 9630 ng/g lipid [19].
Although the wieldy uses of PBDEs were restricted in most developed countries, the high lipophilicity and stability of PBDEs in the environment and their long half-life in human tissues contribute to continuous exposures. For example, an epidemiological study of 1253 women in California suggests that serum PBDE levels continued to increase despite the phase-out [20]. Different health outcomes were found to be affected by PBDEs, with the impairment of male reproduction being among the most sensitive. Human and animal data indicate that PBDEs can affect multiple male reproductive outcomes, such as androgen levels in blood [21], sperm quality [22], and the incidence of genital developmental abnormalities [23]. However, the molecular mechanisms of male reproductive toxicity remain poorly understood [24].

2. Reproductive Health Studies in Humans

Nine epidemiological studies investigated the role of PBDEs on the human male reproductive system. In a prospective cohort study, the effects of ten serum congeners of PBDE (BDE-17, 28, 47, 66, 85, 99, 100, 153, 154, and 183) were cross-sectionally accessed on semen quality in 468 adult men recruited from different counties of Texas and Michigan in the United States. Out of the 10 PBDE congeners studied for five PBDEs, 50% of the serum samples were below LOD, some PBDEs were associated with increased abnormal sperm morphology (BDE-28, 153), and BDE-28 was associated with reduced sperm motility [25].
In a Canadian study of 153 men aged 18–41 years living in Montreal, the hair BDE-47 concentration (median (IQR) 9.4 (4.0–18.0 ng/g) was negatively associated with sperm motility. The quality of the sperm chromatin was not affected by the studied congeners [26]. In Japan, the effect of PBDE was assessed in the sperm samples of 10 young men who were recruited from the Department of Urology at a Private University in Kawasaki. Four congeners of PBDE (BDE-47, 99, 100, and 153) out of the twenty-nine studied were detected in the serum samples. An inverse correlation was observed between the serum levels of BDE-153 (median 0.72 ng/g lipids) and sperm concentration and testis size (r = −0.841, p = 0.002 and r = −0.764, p = 0.01, respectively) [27].
The effects of PBDE on semen quality were assessed in 32 adult men between the ages of 20 and 50 years in a rural community close to an electronic waste recycling area in Qingyuan, China. The sperm parameters were evaluated in relation to concentrations of a range of PBDE congeners in respective house dust and sperm samples. The concentrations of congeners were positively correlated with the dust (median BDE-28 1.46 ng/g, BDE-47 12.0 ng/g, and BDE-153 14.6 ng/g) and paired semen samples (median BDE-28 5.02 pg/g, BDE-47 6.75 pg/g, and BDE-153 7.36 pg/g). Semen BDE-47 was negatively associated with sperm concentration and total sperm count, while the dust levels of BDE-100 were negatively associated with sperm progressive motility and viability [28].
An association between the blood levels of PBDE and male reproductive parameters was studied in men from Ukraine, Poland, and Greenland. Blood and sperm samples from 100 men in each of the countries (total 299) were randomly selected. The effects of BDE-47 and BDE-153 on the men’s reproductive functions were studied using an adjusted linear regression model. BDE-153 and BDE-47 were detected in almost all serum samples, with the highest level in Greenland (median (IQR) for BDE-47 2.0 (0.6–6.9 ng/g lipids) and for BDE-154 2.7 (1.3–7.8 ng/g lipids)), but they were not associated with markers of semen quality and serum reproductive hormone [29].
Studies from a Canadian group reported an association between PBDE concentrations in maternal hair and the incidence of developmental genital malformations, cryptorchidism, and hypospadias in their sons. In a Montreal case-control study, eight congeners of PBDE (BDE-209, 183, 154, 153, 100, 99, 47, and 28) were assessed in the hair of 137 mothers of infants with cryptorchidism and 158 mothers of male infants who did not suffer from cryptorchidism. The concentrations of three PBDEs (BDE-99, 100, and 154) were significantly higher in the cases than in the controls. Double the risk of infants developing cryptorchidism was observed with every ten times increase in the concentration of BDE-99, BDE-100, or BDE-154 in the maternal hair [30]. Similarly, higher levels of total PBDE and five individual congeners (BDE-28, 47, 99, 153, and 154) were found in the hair of mothers whose infants had hypospadias (n = 152) than among controls (n = 64) in Toronto [31][32]. In a multivariable model, hypospadias was associated with a relative 48.2% (95% CI, 23.3–65.4%) higher maternal hair level of total PBDE including eight congeners (Poon et al., 2018). These findings are not supported by the study conducted in California, although the sample size was small [33]. Maternal mid-pregnancy levels of individual PBDE congeners (BDE-28, 47, 99, 100, 153) were not associated with the risk of hypospadias (n = 20 vs. n = 28 control).
In the studies discussed above, PBDEs have been shown to have effects in the general population on various reproductive outcomes in humans, including sperm motility (BDE-28, 47, 100), sperm concentration (BDE-47, 153), total sperm count (BDE-47), sperm morphology (BDE-28, 153), cryptorchidism (BDE-99, 100, 154), hypospadias (total sum of eight PBDEs), and testes size (BDE-153).

3. Reproductive Health Outcomes in Animal Studies

Several studies used animals to study the effects of PBDE on the male reproductive systems.

3.1. Developmental Effects

Many studies focused on the developmental toxicity of BDE-47, the dominating congener in human tissues, using different exposure windows in a rat model. In perinatally exposed rats, many male reproductive outcomes are impaired even if they are assessed many days after exposure cessation. For example, in one study, female rats were exposed to 0.1, 1, and 10 mg/kg/body weight (BW)/day of BDE-47 starting 10 days before mating till lactation cessation on postpartum day 21 [34]. On postnatal day (PND) 88, the pups’ relative testis weight was significantly reduced in rats exposed to all doses compared with the controls. The sperm motility parameters and sperm counts were also significantly reduced in the offspring of dams exposed to 10 mg/kg. In another study, pregnant rats were exposed to 0.2 mg/kg/BW/day of BDE-47 from gestational day (GD) 8 to PND21 [35]. On PND120, the absolute weight of testes, daily sperm production, and percentage of motile sperm were all reduced in the offspring of the exposed dams. In addition, the percentage of morphologically abnormal spermatozoa was higher in the offspring of exposed dams than in the control. This entry also reported significant changes in testes transcriptome, including the suppression of genes essential for spermatogenesis and the activation of immune response genes.
Using the same model, the researchers also demonstrated the developmental effects of BDE-47 on sperm epigenome. First, they showed that perinatal exposure results in an increase in the DNA methylation of epididymal sperm in genes, promoters, and intergenic regions in younger rats (PND65), while the methylation of the same elements decreases in older exposed animals (PND120) [36]. In total, 21 and 9 exposure-related differentially methylated regions (DMRs) were identified in the sperm collected on PND65 and PND120, respectively, with two DMRs overlapping between the two time points. Further, the authors focused on age-dependent changes in the sperm epigenome, namely, profiles of DNA methylation [37] and small non-coding RNA (sncRNA) [38]. They demonstrated that both profiles change significantly with age: 5319 age-dependent DMRs and 1384 sncRNA with age-dependent expression were identified. Perinatal BDE-47 modified the normal dynamics of age-dependent changes in both DNA methylation and sncRNA in sperm. For both epigenetic markers, these changes may be interpreted as an acceleration of age-dependent changes in younger animals and their deceleration in older animals [37][38]. Genes associated with age-dependent DMRs as well as the gene targets of age-dependent sncRNA were highly enriched with categories relevant to embryonic development. Thus, it is plausible that the modification of epigenetic programs in sperm following developmental exposure to PBDE may result in the altered development of embryos conceived by exposed fathers.
Three studies used PBDEs other than BDE-47 to assess the pre-/perinatal male reproductive effects in rats. The perinatal exposure (GD6-PND18) of pregnant/lactating rats to 18 mg/kg/BW of a commercial penta-BDE mix, DE-71 (a mixture of BDE-47, 99, 100, 153, and 154), led to a 1.3-fold increase in pups’ testes weight on PND31 [39]. The in utero exposure to 40 mg/kg/BW DE-71 also resulted in decreased anogenital distance in male offspring, a marker of estrogenic or antiandrogenic effects of exposure [40]. In another study, the male offspring of pregnant rats were exposed to a single dose of BDE-99 at 0.06 or 0.3 mg/kg on GD6; however, no significant changes were observed in the absolute testes and epididymal weight, prostate, seminal vesicles, serum luteinizing hormone, or testosterone levels [41]. The testes of rats exposed to 0.3 mg/kg were considered small when expressed as a percentage of body weight. The daily sperm production, sperm, and spermatid counts were significantly reduced by both doses [41].
Few studies used prepubertal exposure window to study the developmental effects of PBDE in rat models. For example, 14 days of exposure during prepuberty (PND21–PND35) to 0.1, 0.2, and 0.4 mg/kg BDE-47 did not have any effect on the weight of the rats’ testes [42]. An increase in the numbers of Leydig cells, serum testosterone, and decreased serum luteinizing hormone levels were observed upon exposing rats to 0.4 mg/kg/BW. Another study used a prepubertal exposure window (PND23–53) to address the effects of DE-71 in rats. In animals exposed to 3 or 30 mg/kg/BW of DE-71, androgen-dependent tissues (ventral prostate and seminal vesicle), testes, and epididymal weight were not affected. However, 30 mg/kg caused a significant delay in preputial separation by 1.7 days [43].
Two mouse studies investigated the postnatal effects of BDE-209 on male reproductive function. In mice exposed to 10 mg/kg of BDE-209 from PND21 till PND71, no significant difference was observed in the testes, epididymal weight, seminal vesicles, sperm chromatin structure, motility, count, and testes morphology of the exposed group and the control [44]. Conversely, neonatal (PND1–5) exposures to much lower subcutaneous doses (0.025, 0.25, and 2.5 mg/kg/BW/day) resulted in the decreased testicular weight (0.025 and 0.25 mg/kg), epididymal weight (0.25 mg/kg), sperm counts (0.025 mg/kg), elongated spermatid (0.025 mg/kg) and Sertoli cell numbers (0.25 mg/kg) [45]. In addition, a significant reduction in the serum testosterone level was observed after exposure to 0.025 and 0.25 mg/kg of BDE-209 [46].
One recent study analyzed the effects of prenatal (GD1–GD21) exposure to BDE-99 on reproductive development in mice [47]. Doses as low as 0.2 mg/kg/BW decreased the anogenital distance in male offspring, decreased testes size and Leydig cell numbers, increased the incidence of hypospadias, decreased blood testosterone, and resulted in the altered expression of steroidogenic enzymes at the gene and protein levels [47].
Thus, mouse and rat studies showed that perinatal and developmental postnatal exposures to individual PBDE congeners at low, environmentally relevant doses (as low as 0.1 mg/kg/BW/day for BDE-47 and 0.025 mg/kg/BW/day for BDE-209) have morphogenic effects on the male reproductive system, resulting in smaller testis weight and abnormal outcomes of spermatogenesis [34][35][45]. The peripubertal window may be less sensitive to PBDE reproductive toxicity [43][44]; however, more studies are needed in this developmental window. Although several studies demonstrate reduced testes weight following developmental exposures to PBDE, one shows the opposite change [39], suggesting that the direction of change of male reproductive outcomes may depend on the dose, congener, and/or developmental window.

3.2. Adult Effects

Researchers identified five studies that tested the male reproductive toxicity of PBDE in mice and four studies using rat models.
Two mouse studies focused on BDE-47. In one study, 30 days of exposure of adult animals to 0.0015, 0.045, or 30 mg/kg of BDE-47 led to a decrease in the rate of sperm capacitation in all exposed groups [48]. No significant difference between the sperm morphology of exposed and control groups was observed; however, some sperm motility parameters were impaired in some exposed groups. In another BDE-47 study, the histopathological examination of mice exposed to 1.5, 10, and 30 mg/kg of BDE-47 for six weeks showed that sperm decreased in the epididymal lumen at 10 and 30 mg/kg [49].
Other mouse studies analyzed the reproductive effects of BDE-3 and BDE-209. The exposure to 0.0015, 1.5, 10, and 30 mg/kg BDE-3 for six consecutive weeks resulted in decreased sperm counts in mice exposed to 1.5 mg/kg or higher doses, decreased germ cells in seminiferous tubules, and also decreased mature spermatozoa in the epididymis of animals exposed to 30 mg/kg BDE-3 [50]. The exposure of adult mice to 7.5, 25, or 75 mg/kg/BW/day of BDE-209 also resulted in significantly reduced sperm numbers, while only 25 and 75 mg/kg significantly reduced sperm motility [51]. The histopathological examinations also showed a dose-dependent significant reduction in the height of germinal epithelium for the three doses [51]. In another study, adult mice were exposed for 8 weeks to 20, 100, or 500 mg/kg BDE-209 [52].
Two rat studies explored the effects of BDE-47 in adults. Exposures to 0.03 and 1 mg/kg BW/day BDE-47 significantly decreased the serum testosterone levels and increased the number of multinucleated giant cells in the testes that arose from spermatocytes that aborted meiosis, while the daily sperm production was decreased in a group exposed to 1 mg/kg BW/day BDE-47 [53]. In another study, 0.03 and 20 mg/kg BDE-47 also reduced the serum testosterone levels and altered the cellular organization of the seminiferous epithelium. In addition, 20 mg/kg BDE-47 also aborted meiosis in the spermatocytes [54].
Studies using adult rat models also assessed the male reproductive toxicity of BDE-209 and DE-71. In the study with BDE-209, adult male rats were exposed to a range of doses ranging from 1.87 to 60 mg/kg/BW/day for 28 days, and a benchmark dose approach was used to identify the dose that results in a 10% change in the reproductive outcome [55]. For seminal vesicle/coagulation gland weight (most sensitive outcome), a 10% decrease was achieved at the 0.2 mg/kg BW/day exposure level. A dose-dependent decrease of the epididymal weight was observed as well, although the benchmark dose was not calculated for this outcome. No changes in sperm counts or morphology were observed [55]. In adult male rats exposed to 3 and 30 mg/kg DE-71, a decrease in serum androstenedione (3 and 30 mg/kg) and a significant decrease in the ventral prostate (30 mg/kg) were observed, while the serum testosterone (3 mg/kg), LH, and estrone (3 and 30 mg/kg) levels were all increased in the exposed group [56].
Few field studies demonstrated the effects of PBDEs in large mammals. For example, the testicular levels of PBDEs negatively correlated with Sertoli cell numbers and germ cells’ proliferative activity in dogs residing in Finland, Denmark, and the UK [57]. Testis size was significantly negatively associated with subcutaneous levels of PBDE in 20 subadult polar bears in Greenland [58].
Rodent laboratory studies as well as wildlife studies demonstrate that PBDE exposure is a potent factor affecting male reproductive system physiology in adult age.

4. Most Sensitive Outcomes

The most sensitive male reproductive outcomes identified in developmental and adult animal experiments are as follows. Among the developmental studies, the lowest dose capable of producing health effects was reported in the study where mice were exposed to 0.025 mg/kg/BW/day of BDE-209 for five days following birth [45]. This dose of BDE-209 decreased the testicular weight and the sperm and elongated spermatid count. The subcutaneous delivery of BDE-209 used has low relevance for human exposure, as humans are mostly exposed to PBDE via inhalation and ingestion.
Among adult studies, the lowest dose that produced changes in male reproductive outcomes was 0.0015 mg/kg of BDE-47 per day over 30 days [48]. That dose resulted in decreased sperm motility and sperm capacitation in rats. Thus, in adult animals, PBDEs may affect the physiology of spermatozoa at doses at which no effects on reproductive organ size or sperm production are seen.
In both developmental and adult rodent studies, the effects of PBDEs on the male reproductive system were seen at a microgram range of exposures, suggesting that in developed countries, almost 100% of males may experience changes in reproductive functions due to exposure to PBDE during their lifespan.

5. Mechanisms That Mediate Male Reproductive Toxicity of PBDE

Several different molecular and endocrine mechanisms that may be involved in the reproductive toxicity of PBDEs have been suggested. These mechanisms are reviewed below.

5.1. Induction of Oxidative Stress

Oxidative stress is caused by the inability of the body to manage the imbalance in the cellular production and deactivation of reactive oxygen species (ROS) [59]. The oxygen-containing reactive species include superoxide radicals (O2), hydroxyl radicals (OH), hydrogen peroxide (H2O2), peroxyl radical (LOO), and lipid hydroperoxides (LOOH) [60][61]. ROS are natural by-products of oxygen metabolism, and they participate in several physiological processes such as the immune response to pathogens, protein phosphorylation, and cellular signaling [59]. Spermatozoa are a source of ROS themselves and, additionally, ROS are produced in semen in large quantities by polymorphonuclear leucocytes [62][63][64][65].
The sperm membrane of spermatozoa has an unusual structure, which is enriched with lipids that are the main substrates for peroxidation: phospholipids, sterols, and saturated and polyunsaturated fatty acids [66][67][68]. This membrane structure makes spermatozoa particularly susceptible to damage by ROS [69][70][71][72]. Thus, excessive ROS can significantly affect sperm physiology through their effect on the sperm plasma membrane, resulting in decreased sperm motility and vitality [73][74][75][76], impaired capacitation [77], acrosome reaction [78], and other sperm parameters and outcomes of fertilization [79]. Although spermatozoa have an antioxidant defense system that detoxifies ROS, the overproduction of ROS can overwhelm these mechanisms and induce oxidative stress [80]. Overall, oxidative stress is one of the major causes of sperm damage involved in male infertility [80][81][82][83].
The exposure to xenobiotics including PBDEs can increase ROS production and induce oxidative stress in different tissues [84][85][86]. A growing body of studies using diverse biological systems demonstrate that PBDEs can affect ROS balance and induce changes to the antioxidant defense system. To name a few, the exposure of GC-2 cells to 8, and 32 μg/mL of BDE-209 led to an increase in ROS levels in the cell [51]. Metabolites of BDE-47 (3-MeO-BDE47, 3-MeO-BDE47, 5-MeO-BDE47, and 5-OH-BDE47) increased the activity of superoxide dismutase (SOD), decreased the levels of glutathione (GSH), and increased ROS in LO2 cells in a dose-dependent manner [87]. In another study, the exposure of LO2 cells to 10 and 50 μM BDE-209 also led to a significant increase in ROS levels [88]. The treatment of HS-68 human cell culture to 50 μmol/L BDE-47, 100 μmol/L BDE-99, and 2 μmol/L BDE-209 led to an increase in the levels of intracellular ROS [86]. In comparison with the control, a significant increase in the superoxide dismutase enzyme activity and a significant decrease in GSH reductase enzyme activity was observed in the erythrocytes of adult rats exposed to 0.6 and 1.2 mg/kg of BDE-99 [89].
Thus, it is not surprising that oxidative stress is considered a candidate mechanism that may connect PBDE exposure with adverse male reproductive outcomes. Several in vivo studies addressed this potential mechanism of reproductive toxicity mostly focusing on BDE-209. Testicular malondialdehyde (MDA), a marker of lipid peroxidation, was increased in mice testes exposed to 25 and 75 mg/kg of BDE-209, while SOD enzymatic activities were decreased [51]. In comparison with the control, SOD activity and GSH levels were also decreased in male mice dosed with 200 and 500 mg/kg/BW of BDE-209 [90]. MDA was also increased in adult mice testes exposed to 100 and 500 mg/kg/BW BDE-209 for 8 weeks, but not in mice exposed to 20 mg/kg [52]. In the same study, GSH levels were increased in the 20 and 500 mg/kg groups, although small n (4–5) and very high SD in the 20 mg/kg group may indicate a false positive finding. On the other hand, the increased expression of gene markers of endoplasmic reticulum stress (Atf6, Ire1) and gene (Bcl-2) and protein (Bax) markers of apoptosis in the 20 mg/kg group may indicate that this level and duration of exposure is sufficient to reach an oxidative stress response [52]. A significant increase in the sperm H2O2 was observed in mice exposed to 500 and 1500 mg/kg of BDE-209 [44]. In another study, mice were exposed to 10, 500, and 1500 mg/kg/BW of BDE-209 from GD 0 to 17 [91]. A significant increase in sperm H2O2 generation was observed at the lowest and highest doses [91]. In lactating female mice orally gavaged with 500 and 700 mg/kg/BW of BDE-209, SOD and catalase enzyme activities were decreased in male pups on PND21 and PND28 from both exposure groups [92]. Moreover, the maternal exposure of female mice to 500 and 700 mg/kg of BDE-209 led to a decrease in the activities of catalase and SOD enzymes in the testes of the offspring. In the same study, an increase in the total ROS production was observed in the testes of offspring from all exposure groups [93].

5.2. Metabolic Disruption

Recent research demonstrated that PBDEs are potent metabolic disruptors affecting lipid and glucose metabolism at environmentally relevant doses, although the mechanisms of this disruption are poorly understood. For example, serum triglycerides were significantly increased two-fold in mouse pups exposed to 0.2 mg/kg of BDE-47 perinatally [94]. Neonatal exposure to 1 mg/kg of BDE-47 produced the opposite effects on lipids in circulation: decreased blood triglycerides and increased liver triglycerides in adult mice [95]. In another study, perinatal exposure to 0.002 and 0.2 mg/kg BDE-47 caused a worsening of high-fat diet-induced obesity, hepatic steatosis, and impaired glucose homeostasis [96]. A recent mouse study reported the development of a diabetic phenotype in mice perinatally exposed to 0.1 mg/kg BW DE-71 [97]. These mice developed fasting hyperglycemia, glucose intolerance, and reduced insulin sensitivity. In rats exposed to 14 mg/kg BW/day of a mix of PBDE congeners, a decrease in insulin-stimulated glucose oxidation and an increase in isoproterenol-stimulated lipolysis was observed in adipocytes [98]. A significant disruption of the metabolism of lipids and carbohydrates was also observed in mice exposed to high doses of BDE-209 [88][92].
Similar evidence started to emerge recently from epidemiological studies. For example, among 34 obese Qatari individuals, insulin-resistant subjects had significantly higher levels of BDE-99, BDE-28, BDE-47, and the sum of penta-BDE in adipose tissue than insulin-sensitive counterparts [99]. Additionally, BDE-99 and BDE-28 positively correlated with fasting insulin levels. An increased risk of type 2 diabetes (T2D) in association with dietary exposures to PBDE was found in a French study of 71,415 women including 3667 diagnosed with T2D, although the exposure assessment was conducted using a scenario evaluation approach based on a questionnaire and may suffer from low accuracy [100]. Similarly, a positive association was found between total PBDEs and gestation diabetes mellitus (GDM) in Tehran women (70 cases vs. 70 controls) [101]. Another study on 147 mother–children pairs demonstrated that in utero BDE-99 was associated with lower childhood levels of triglycerides, high-density lipoprotein, and total lipids in children’s blood at 6–7 years of age [102], suggesting long-lasting metabolic effects of early-life exposure.
Overall, experimental and epidemiological evidence suggests that PBDEs may mimic and exacerbate the effects of a high-fat or high-calorie diet as they stimulate lipid accumulation in the liver and decrease glucose tolerance and insulin sensitivity. These symptoms resemble diabetic and obesity phenotypes, both contributing to male infertility via multiple mechanisms [103][104][105][106]. A discussion of these mechanisms is beyond the scope of the current entry. However, it is important to mention that the metabolic disruption associated with obesity and/or diabetes results in a negative impact on semen parameters, including sperm concentration, motility, viability, and normal morphology [103][104][105][106].
Evidence that supports the hypothesis that PBDE reproductive toxicity is mediated via metabolic disruption was obtained recently in a metabolomics study where 76 differential metabolites were found in mouse testis tissue following exposure to 0.0015, 1.5, 10, or 30 mg/kg/BW of BDE-3 for six weeks [50]. These metabolites were enriched for nucleotide metabolism and lipid metabolism as well as pathways involved in the metabolism of several amino acids and riboflavin, an important player in carbohydrate energy metabolism.

5.3. Inflammatory Response

Emerging evidence indicates that PBDEs may promote inflammation in mammalian tissues. For example, positive relationships between PBDEs and pro-inflammatory cytokines (IL-6 and TNF-α) in circulation were reported in pregnant and postpartum women [107]. Studies demonstrated that PBDEs modulate inflammatory pathways in the human placenta [108][109][110]. Proinflammatory cytokines have been associated with increased ROS production, a decrease in the production of testosterone in Leydig cells, and a decrease in sperm motility and concentration [111][112][113]. The role of PBDE-induced inflammation on male reproductive outcomes is not yet well understood.
For instance, inflammatory cell infiltration was observed in the epididymis interstitium of mice dosed with 30 mg/kg/BW of BDE-3 [50]. In another study, the suppurative inflammation of the epididymis in mice was observed after exposure to 30 mg/kg of BDE-47 for 6 weeks [49]. Additionally, the proteomic and metabolomic analysis of testis tissue showed that BDE-47 can trigger the apoptosis and inflammatory pathway [49]. However, this analysis was carried out for a merged list of proteins and metabolites affected by either dose used in the study: 1.5, 10, and 30 mg/kg/day daily for 6 weeks. Given that two higher doses have very low relevance to real-life human exposure, researchers conducted Metascape enrichment analysis [114] for the list of proteins affected by the low dose, which resulted in no enriched biological categories relevant to inflammation.
An increase in the expression of inflammatory response genes was observed in the testes of adult rats after perinatal exposure to 0.2 mg/kg of BDE-47 [35] and in the testes of immature rats after prenatal exposure to 0.2 mg/kg of BDE-99 [47]. The biological categories enriched with upregulated genes included interferon, IL3 and IL5, and TNF-α signaling, allograft rejection, and natural killer cell-mediated cytotoxicity, among others. Additionally, exposed animals had significantly smaller testes, decreased sperm production, and an increased percentage of morphologically abnormal spermatozoa [35]. Similar changes in male reproductive function (decreased testis weight and an increased percent of morphologically abnormal spermatozoa) were reported in adult mice exposed to inflammatory challenge [115]. Taken together, these findings suggest that PBDE-induced changes in immune response genes in the testes may be causally linked with the disruption of male reproductive function.

5.4. Disruption of Blood–Testis Barrier (BTB)

A growing body of literature provides clear evidence that early-life exposure to PBDEs at environmentally relevant doses may have long-lasting effects on mammalian tissue physiology [94][95][116][117], including male reproductive tissues [34][35][42][45][47]. Studies demonstrate that perinatal or neonatal exposures result in reduced testis weight and reduced sperm quality parameters [34][35][45][47][49].
One study demonstrated that exposure to BDE-209 may result in decreased activity of cortactin (CTTN) due to decreased CTTN expression and the increased Tyr phosphorylation of CTTN [45]. Upon activation, CTTN recruits Arp2/3 complex proteins to actin microfilaments, inducing actin branching from nucleation sites [118]. Thus, decreased CTTN activity in response to BDE-209 may disrupt ectoplasmic specialization [119][120], an actin-based junctional structure between the Sertoli cells and Sertoli cells (basal endoplasmic specialization) or Sertoli cells and germ cells (apical endoplasmic specialization) [121]. Ectoplasmic specialization is an essential mechanism supporting the blood–testis barrier (BTB), a structure that regulates the biochemical environment in the apical compartment of the seminiferous epithelium in which germ cells develop and protects germ cells from toxic compounds and autoimmune response [122]. Intact BTB is necessary to avoid the production of anti-sperm antibodies and autoimmune response leading to male infertility [123]. Damage of the BTB is associated with inflammation, germ cell loss, reduced sperm count, and ultimately subfertility or infertility [124]. From this point of view, it is interesting that in adult animals developmentally exposed to PBDE, inflammatory response genes were upregulated in the testes [35]. Similarly, the immune response pathways were enriched in the testes of immature rats prenatally exposed to 0.2 mg/kg body weight or higher doses of BDE-99 [47].
Thus, one hypothesis may explain the developmental male reproductive toxicity of PBDEs via their effects on the BTB formation and subsequent chronic inflammation caused by “leaky” BTB. One recent study report decreased the expression of tight junction proteins ZO-1 and β-catenin in mice testes following 8 weeks of adult exposure to 20 mg/kg/BW or higher doses of BDE-209 [52]. At higher doses (100 and 500 mg/kg/BW), the discontinued structure of the BTB junctions was observed. Additionally, one study indirectly supports the BTB hypothesis, demonstrating that the gestational and lactational exposure of rats to 0.06 mg/kg/day of a mix of brominated flame retardants (technical PBDE mixtures DE-71, DE-79, and BDE-209, and HBCDD) significantly downregulated adherens junction proteins, E-cadherin, and β-catenin, and the gap junction protein connexin 43 (Cx43) in post-pubertal mammary glands [125]. These proteins play important roles in the BTB integrity, and changes similar to these observed in the mammary gland may cause BTB disruption in testes.

5.5. Endocrine Disruption: Testosterone Signaling

PBDEs are well-recognized endocrine disruptive chemicals (EDCs). Specifically, a substantial body of literature connects PBDE exposures with reproductive hormone signaling.
PBDEs may have opposite effects on testosterone synthesis during peripubertal and adult exposures. For example, the exposure of rats to environmentally relevant doses of BDE-47 during the peripubertal window (PND21–35) resulted in Leydig cell hyperplasia, the increased expression of steroidogenesis enzymes in Leydig cells, and, ultimately, increased serum testosterone [42]. In another study, Leydig cells were obtained from prepubertal rats (PND49) and used to analyze the effects of BDE-47 in vitro. The testosterone production was three-fold higher at a 1 mM BDE-47 concentration compared with the control [126]. Similarly, in humans, developmental exposure to PBDE may increase testosterone secretion later in life. For example, in a birth cohort study, a 10-fold increase in maternal prenatal serum concentrations of BDE-153 was associated with an approximately 92.4% increase in testosterone in 12-year-old sons [127]. Not all studies support the increased production of testosterone following developmental exposure to PBDE. For example, in young prepubertal rats, the serum testosterone level significantly decreased following prenatal exposure to 0.2 mg/kg body weight BDE-99 and higher doses [47]. Similarly, 35-day-old mice had decreased blood testosterone and decreased numbers of Leydig cells following prenatal exposure to 0.2 mg/kg/BW and higher doses of BDE-99 [47].
In experiments where adult rats were exposed to low doses of BDE-47, the blood testosterone levels were significantly decreased in a dose-dependent manner [53][54]. Blood testosterone was also decreased in adult mice exposed to high doses of BDE-209 [51][128]. Decreased testosterone was also seen in mice exposed via mother’s milk to BDE-209, although the relevance of these findings is low due to the use of very high doses (500 and 700 mg/kg) [93]. Human data on testosterone levels in relation to adult exposures have only started to emerge. For example, in one study of 63 US men, positive associations of octa-BDE in house dust with serum testosterone and an inverse association of deca-BDE in house dust with testosterone were reported [129]. BDE-47 was also positively associated with testosterone levels in adult male sport fish consumers [21].
Many studies directly tested the ability of PBDEs to bind to androgen receptors (AR) using ex vivo, in vitro, and in silico experiments. Some of these studies analyzed the ability of individual PBDE congeners, their mixes, and their hydroxylated and methoxylated metabolites to interfere with androgen receptors using in vitro assays in which reporter cell lines carry a luciferase gene under the transcriptional control of response elements for activated AR (chemically activated luciferase gene expression (CALUX) assay) [56][130][131]. These studies did not identify any PBDE congener or metabolite with AR-agonistic activity [130][131]. However, they did report that many PBDEs have AR-antagonistic activity at physiologically relevant concentrations.
For example, one study of 19 PBDE congeners reported four congeners (BDE-19, BDE-100, BDE-47, and BDE-49) with IC50 values lower than IC50 for flutamide, a reference AR-antagonistic compound and nonsteroidal antiandrogenic drug [130]. Only three congeners did not show AR antagonistic activity in that study. Similarly, BDE-47, BDE-100, and industrial penta-BDE mixture DE-71 inhibited dihydrotestosterone (DHT)-induced AR activation by 50% at 5 μM concentrations in another study, while BDE-99, BDE-153, and BDE-154 did not show AR antagonism [56].
Another study reported the AR antagonistic activity of 12 out of 16 tested PBDE congeners and metabolites with IC20 in a range of 10 nM–1 μM [131]. The authors reported that the anti-androgenic activities of 4′-HO-BDE-17 and BDE-100 were about 5- and 10-fold lower than that of hydroxyflutamide, respectively. Specifically, the IC20 for 4′-HO-BDE-17 was determined to be as low as 0.086 μM. Similar experiments conducted by different research groups resulted in higher values for inhibitory concentrations of hydroxylated metabolites of BDE-47 [132]. In that study, the IC50 of 4′-HO-BDE-17 was 1.41 μM, and 6-HO-BDE-47 was identified as the most potent hydroxylated metabolite (IC50 = 0.34 μM). In a study of methoxylated PBDE metabolites, 6-MeO-BDE-47 was identified as a potent anti-androgen with IC50 = 41.8 μM [133].
The AR-antagonistic activity of PBDEs was also assessed ex vivo in a competitive binding assay with rat ventral prostate cytosol [56]. DE-71 and BDE-100 inhibited the AR binding of a radiolabeled synthetic androgenic steroid [3H]R1881, with an IC50 of approximately 5 μM [56]. An in silico study using an induced fitting dock test and binding affinity estimation showed that BDE-47, BDE-99, and their methoxylated and hydroxylated metabolites are tightly bound to the AR binding site in similar pattern to how testosterone binds to AR [134]. The methoxylated metabolites of BDE-47 and BDE-99 had higher binding energy than the parent compounds, and 6-MeO-BDE-99 had the highest binding energy among all tested metabolites, close to the binding energy of testosterone.
In the general population, PBDE concentrations in blood may reach the nM range [102][135][136], suggesting that in highly exposed individuals, PBDE and their metabolites may disrupt androgenic signaling by the direct antagonistic interaction with AR. Additionally, changes in testosterone signaling may result from its decreased production and/or altered hypothalamic–pituitary control.

5.6. Endocrine Disruption: Estrogen Signaling

PBDE also demonstrates xenoestrogenic properties in a range of in vitro studies. Animal studies that provide unambiguous support of (anti)estrogenic properties of PBDEs were not able to be identified. In the only identified human study, maternal serum BDE-145 at 35 weeks of pregnancy in a Dutch cohort was positively correlated with estradiol (E2) and free E2 levels in their sons at 3 months of age [137].
Several studies used estrogen receptor (ER)-CALUX assays in different cell lines to assess (anti)estrogenic properties of PBDE and their metabolites. The assessment of 17 PBDE congeners showed that 11 compounds have agonistic activity with concentrations leading to a 50% induction (EC50) ranging from 2.5 to 7.3 μM [138]. The potencies of these congeners were 250,000–390,000 times lower than the potency of the natural ligand, estradiol (E2). An ER-CALUX assay in the presence of E2 demonstrated that some hexa-BDE and one hepta-BDE have ER-antagonistic properties at IC50 in the 0.8–3.1 μM range. In an ERα-specific assay, 4′-OH-BDE-30 demonstrated estrogenic properties with an EC50 < 0.1 μM (four orders of magnitude lower than EC50 for E2) [138]. Another study tested the estrogenic activity of 12 PBDE congeners and their hydroxylated metabolites using ER-CALUX assay in MCF-7 cells [139]. Two hydroxylated metabolites, 4′-OH-BDE-17 and 3′-OH-BDE-7, but not other tested compounds, exhibited estrogenic activity at 1–10 μM concentrations. The estrogenic activity of OH-PBDE was also tested in vitro in two luciferase reporter gene systems in another study [140]. The EC50 of the most potent metabolite, 4′-OH-BDE-17, was determined as 4.7 μM, while the EC50 of E2 was 1.2 pM [140]. Another ER-CALUX study in MCF-7 cells 10 of the 22 OH-PBDEs exhibited luciferase induction at μM concentrations and demonstrated 105- to 107-fold smaller potency than E2 [141]. Additionally, six HO-PBDEs showed ER-agonistic effects, four showed no effects, and twelve demonstrated ER-antagonistic effects. Interestingly, the ER agonists in these studies had four or fewer bromine atoms, while the ER antagonists had four or more bromine atoms [141].
ER-agonistic properties were also found for low-brominated PBDE, up to hexa-brominated BDE-155 in a study of 19 PBDE congeners and 6-OH-BDE-47 using ER-CALUX assay in T47D human breast cancer cells [130]. The EC50 values were >2 μM, six orders of magnitude lower than for E2. Tetra-congener (BDE-79), all tested hepta-congeners (BDE-181, BDE-183, BDE-185, and BDE-190), and 6OH-BDE-47 demonstrated ER-antagonistic properties. 6-OH-BDE-47 was the most potent anti-estrogen, with IC50 = 0.5 μM, which is around 3000 times lower potency than that of a reference antiestrogenic drug ICI 182.780 [130]. The ER-agonistic properties of low-brominated PBDE and ER-antagonistic properties of higher-brominated PBDE were confirmed in another study carried out by the same research group [142]. In a study of eight PBDE congeners and their metabolites in ERα and ERβ CALUX assays in BG1Luc4E2 ovarian cancer cells, three (BDE-28, BDE-47, and BDE-100) showed ERα-agonistic properties, BDE-100 also showed ERβ-antagonistic properties, and two other congeners (BDE-99 and BDE-153) showed antagonistic properties for both ER receptors in the μM range [131]. Among the metabolites, 4′-HO-BDE-17 showed the most potent estrogenic activity (EC50 = 0.2 μM for both ERα and ERβ assays), and 4′-HO-BDE-49 showed the most potent anti-estrogenic properties (EC50 = 2.3 and 3.6 μM for ERα and ERβ assays, respectively) [131]. Among the BDE-47 hydroxylated analogs, 4′-HO-BDE17 induced a significant estrogenic response, while other compounds showed anti-estrogenic potency (4′-HO- BDE-17, 6-HO-BDE-47, 2′-HO-BDE-28, BDE-47) or no (anti)estrogenic activity (4′-HO-BDE-49) in the micromolar range [132].
The direct binding of 22 OH-PBDEs with the human ERα ligand-binding domain was measured using a surface plasmon resonance technique using E2 as the positive control [141]. The KD (equilibrium dissociation constant) of E2 was 0.35 nM. Seven out of the twenty-two OH-PBDEs showed a direct binding reaction at the ERα ligand-binding domain with KD values in the range of 0.15–7.90 μM. The relative binding potency of the most potent compound, 6-OH-BDE-47, was 0.24% of E2. Six OH-PBDEs, metabolites of DE-71, were tested in a competitive binding assay with recombinant ERα and tritiated E2 (3H-E2.) [140]. All of the OH-PBDE displaced 3H-E2 from ERα, but their binding affinities in relation to E2 ranged from 0.001% to 0.03%. 4′-OH-BDE-17 and 4′-OH-BDE-49 were the most potent metabolites, with IC50 in the micromolar range.
Xenobiotics may induce (anti)estrogenic singling via both canonical nuclear estrogen receptors (ERs) as well as via nongenomic G protein-coupled estrogen receptor (GPER) pathways [143]. The ability of PBDE to induce nongenomic estrogenic signaling was assessed in a recent study where the binding affinities of 12 PBDE congeners and their 18 hydroxylates metabolites to GPER were determined in a competitive binding assay in vitro. Eleven hydroxylated PBDEs, but none of the PBDEs, bound to GPER directly with EC50 ranging from 1.3 to 20 μM and relative binding affinities ranging from 1.3% to 20.0% compared to E2 [144].
According to another hypothesis, PBDEs may exert estrogenic properties via a non-ER-dependent mechanism via the inhibition of estradiol sulfotransferase enzymes [145]. The decreased sulfation of E2 may result in an increased bioavailability of natural estrogens and increased estrogenic signaling. The ability of a recombinant human sulfotransferase 1E1 (SULT1E1) to produce E2 sulfate from tritiated E2 in the presence of PBDE congeners and their hydroxylated metabolites were tested ex vivo [130][142][145]. Most PBDE congeners inhibited SULT1E1 at μM concentrations, while OH-PBDEs had IC50 in a range of 0.2–1.4 μM.
The studies discussed in this section demonstrate that (1) low-brominated PBDEs and their metabolites mostly show estrogenic activity, while highly brominated compounds demonstrate anti-estrogenic properties; (2) hydroxylation mostly increases the estrogenic or anti-estrogenic properties of PBDE; (3); PBDEs and their metabolites may interact with estrogenic signaling via a different mechanism at the μM range; and (4) PBDEs and their hydroxylated metabolites have activities towards ERs 104 to 107-fold less potent than E2. These findings, taken together with the lack of populational studies indicating changes in circulating estrogens in relation to PBDEs, suggest that the disruption of estrogenic signaling is an unlikely mechanism of the male reproductive toxicity of PBDEs.

5.7. Endocrine Disruption: Luteinizing Hormone (LH) Signaling

Human studies indicate that early-life exposure to PBDEs results in increased LH secretion later in life. For example, in a birth cohort study, a 10-fold increase in maternal prenatal serum concentrations of BDE-100 and BDE-153 were associated with approximately 75% and 97% increases in LH in 12-year-old boys, respectively [127]. Similarly, the amount of PBDEs in maternal milk positively correlated (p < 0.033) with serum LH levels in three-month-old infants in a prospective Danish–Finnish study [146]. Additionally, serum LH levels correlated positively with individual congeners BDE-47, BDE-100, and BDE-154. A rat study demonstrated the opposite relationship between developmental PBDEs and LH production [42], although the study used a different exposure window (prepubertal) [42].
Studies of LH association with adult exposures produce controversial results. For example, in a study of 24 men recruited through a US infertility clinic, the concentrations of BDE-47, BDE-99, and BDE-100 in house dust were inversely associated with serum LH [147]. However, in an extended study of 63 participants by the same group, these associations were not significant [129]. Additionally, the study reports positive associations between the dust concentrations of octa-BDEs and serum LH [129]. In a study of 27 adult US men, LH serum concentrations were positively associated with BDE-47 and BDE-99; however, the positive relationships were almost absent following the removal of a single influential data point [148]. Similarly, no association between PBDEs and LH was observed in 77 men working in e-recycling facilities or other recycling facilities (low-exposure group) [149]. In another e-recycling facility study of 76 men in southern China, a positive correlation was found between semen or serum levels of BDE-153 and serum LH [150].
Several rodent studies analyzed LH response to PBDE exposures. For example, blood LH was also increased in adult male rats exposed to 3 and 30 mg/kg DE-71 [56] and in 35-day-old mice exposed perinatally to 20 mg/kg/BW of BDE-99 [47].

5.8. Endocrine Disruption: Follicle-Stimulating Hormone (FSH) Signaling

There are only a few studies that report FSH levels in response to PBDE exposure in population studies, and their findings are conflicting. In a birth cohort study, a 10-fold increase in maternal prenatal serum concentrations of BDE-153 was associated with approximately 22% increases in FSH in 12-year-old boys, respectively [127]. Other studies analyzed FSH in response to PBDE in adult subjects. In a study of 24 men recruited through a US infertility clinic, the concentrations of PBDE in house dust were inversely associated with serum FSH [147]. A reanalysis in an extended study of 63 participants demonstrated a significant inverse association between the dust concentrations of penta-BDEs and serum FSH [129]. In a southern China study of 54 highly exposed and 58 moderately exposed participants, negative associations between serum FSH and serum concentrations of several PBDE congeners were found in female, but not male, participants [151]. Another study of 27 USA men reported the opposite relationship between PBDE and FSH: BDE-47, BDE-100, and BDE-153 were significantly positively associated with FSH among older men (≥40 years old), but not younger men (<40 years old) [148]. No association between PBDEs and FSH was observed in 77 men working in e-recycling facilities or other recycling facilities (low-exposure group) in southern China [149]. One recent study report increases of blood FSH in 35-day-old mice following prenatal exposure to 20 mg/kg/BW of BDE-99 [47].

5.9. Endocrine Disruption: Inhibin-B and Sex Hormone-Binding Globulin (SHBG)

There are only a few studies that report inhibin B and/or SHBG levels in response to PBDE exposure in population studies. The amount of PBDEs in maternal milk did not correlate with the serum levels of inhibin B and SHBG in three-month-old infants in a prospective Danish–Finnish study [146]. However, in another study from the Netherlands, maternal serum BDE-145 at 35 weeks of pregnancy positively correlated with inhibin B levels, but not SHBG, in their sons at 3 months of age [137].
In a study of 24 men recruited through a US infertility clinic, concentrations of BDE-47, BDE-99, and BDE-100 in house dust were positively associated with inhibin B and SHBG [147]. SHBG was also positively associated with penta-BDE in an extended study of 63 participants, while the association of inhibin B with PBDEs became non-significant [129]. In a study of 27 USA men, serum PBDEs were inversely associated with inhibin B [148].

5.10. Thyroid Hormone Signaling

Thyroid hormones play an important role in male reproductive health. Hypothyroidism is associated with impaired concentrations of reproductive hormones and abnormalities in sperm morphology, while thyrotoxicosis induces abnormalities in sperm motility [152][153]. Thyroid hormones (TH) inhibit Sertoli cell proliferation [154] and promote their differentiation [152]. They also suppress the tight junctions between Sertoli cells and spermatogonia [155]. Additionally, thyroxine (T4) has a direct effect on sperm motility [156]. Overall thyroid dysfunction may result in reduced fertility and infertility [152]. Altogether, this evidence suggests that thyroid disruption by PBDE may have multifaceted effects on male reproduction.
Due to the structural similarity of PBDEs and thyroid hormones, the ability of PBDEs to affect thyroid signaling has attracted a lot of attention from the research community. A substantial body of evidence exists to date demonstrating thyroid disruption by PBDEs in the general population and in laboratory animals at environmentally relevant doses. For example, the pre- or perinatal exposure to low doses of BDE-47 resulted in significant decreases in T3 and T4 in rats [34][157] and sheep [158]. Additionally, T3 was negatively correlated with sperm counts [34].
Human population studies produce controversial results on the associations between circulating THs and PBDE exposure. For example, a significant correlation between serum PBDE and thyroid dysfunction was documented in a recent population study in China [159]. A recent meta-analysis of 16 population studies concluded that serum THs were negatively associated with serum PBDEs when the median levels of PBDEs were <30 ng/g lipid; there was no correlation between THs and PBDEs at median levels between 30 ng/g and 100 ng/g lipids, and the associations were mostly positive if the median levels of PBDEs were >100 ng/g lipids [160]. These potentially U-shaped dose–response relations within the range of human exposure in the general population make the analysis of the role of thyroid disruption by PBDEs on male reproductive health particularly complex.
PBDE can likely affect thyroid signaling via several molecular mechanisms. The most well-documented mechanism consists of TH displacement by PBDEs and their metabolites from TH transport proteins. For example, the binding constants of 14 OH-PBDEs with transthyretin (TTR) and thyroxine-binding globulin (TBG) assessed by competitive fluorescence displacement assay were in a range of 69–140 nM for TTR and 22 nM–6.5 μM for TBG, with tetrabrominated compounds having the highest binding affinity [161]. The dissociation constant (Kd) for TTR and 11 OH-PBDEs were in the nM range in another study using different methodological approaches to analyze the binding affinities of PBDE with TH transport proteins. [162]. The Kd of 10 and 5 OH-PBDE were lower than the Kd of T3 and T4, respectively. For TBG, the Kd of 7 OH-PBDE was also in the nM range and comparable with the Kd of THs. Other studies support these findings [130][163][164]. Additionally, PBDEs and their metabolites may affect other aspects of thyroid signaling, including the hypothalamic–pituitary control of thyroid signaling, the conversion of T4 to T3 by deiodinases in tissues, and the interaction with multiple nuclear receptor isoforms [165][166][167].
The well-documented ability of PBDE and their metabolites to disrupt thyroid signaling in the general population together with experimental toxicological evidence and mechanistic studies demonstrate that thyroid disruption is a plausible mechanism of the male reproductive toxicity of PBDEs. The detailed discussion of the mechanisms of thyroid disruption by PBDEs is outside the scope of the current study.

5.11. Insulin-like Growth Factor (IGF) and Mechanistic Target of Rapamycin (mTOR)

IGF-1 is another important metabolic hormone that is poorly studied both in relation to men’s reproductive health and as a target of endocrine disruption by PBDEs. Understanding of the role of the growth hormone (GH)/IGF system in human reproductive physiology started to emerge only recently [168]. IGFs participate in sexual differentiation during fetal development and promote puberty onset [169]. Lower IGF-I levels are associated with impaired sperm parameters [170].
Two studies done in rats and mice demonstrated that perinatal exposures to low, environmentally relevant doses of BDE-47 result in a long-lasting increase in plasma IGF-1 [94][171]. Interestingly, in a mouse study, a two-fold increase in the circulating IGF-1 was observed in males on PND140, following perinatal exposure to 0.2 mg/kg of BDE-47 [94], suggesting permanent changes in IGF-1 signaling following early life programming.
Human data are scarce. For example, a positive correlation was observed between umbilical cord blood PBDEs levels and placental expression of insulin-like growth factor-binding protein 3 (IGFBP-3), the major IGF-1 transport protein in the blood [172]. In addition, a significant positive correlation was observed between BDE-154, BDE-209, and IGF-1 mRNA levels in the placenta. A cord blood log of IGF-1 levels was also positively correlated with the log of BDE-196 level in breast milk and negatively with the log of BDE-85 in breast milk [173].
The binding of IGF-1 to its respective receptor IGF-1R triggers a PI3K/Akt cascade, resulting in the activation of the mechanistic target of rapamycin (mTOR) complex one (mTORC1) and two (mTORC2). Therefore, the ability of PBDEs to induce mTOR shown in studies discussed below likely provides additional evidence of IGF-1 disruption.
The mTOR-centered pathway is a metabolic master-switch, which, at starvation, suppresses biosynthetic programs and increases the recycling of proteins and organelles to provide an internal resource of metabolites [174]. Conversely, the stimulation of the pathway by nutrients and growth factors causes the activation of biosynthesis and the suppression of autophagy [175]. According to the human protein atlas, the highest levels of mTOR RNA expression are found in the testes and the highest levels of mTOR protein expression are found in the prostate in comparison with all other human tissues [176]. Patients using mTORC1 inhibitors for immunosuppression therapy experience spermatogenesis disruption [177][178][179][180].
In the testes, the mTOR pathway plays multiple roles. One important tissue-specific role of the pathway consists in the regulation of the BTB integrity. Recent studies have shown that BTB integrity is determined by the balance between the activities of the two mTOR complexes, with mTORC1 promoting disassembly of the BTB and mTORC2 promoting its integrity [181][182]. The inhibition of mTORC1 by rapamycin resulted in BTB strengthening [183]. It was demonstrated further that ribosomal protein S6 (rpS6), the downstream phosphorylation target of mTORC1, is essential for F-actin network restructuring in Sertoli cells [183]. The same research group demonstrated that the knockdown of Rictor (a key component of mTORC2) in Sertoli cells in vitro was associated with a loss of barrier function, changes in F-actin organization, and loss of interaction between actin and adhesion proteins [184].
Published transcriptomic changes in response to PBDE exposure identified changes in ribosomal genes as a molecular signature of exposure [95]. Given that ribosomal genes are controlled by the mTOR pathway [185][186], their coordinated changes may be used as a surrogate measure of mTOR activity. It was shown that 0.2 mg/kg body weight of BDE-47 increased the expression of ribosomal genes in different species (mice, rats) and different tissues (liver, brain frontal lobes). However, the effects of higher doses of BDE-47 or DE-71 were the opposite and corresponded to mTOR suppression [95].
Overall, the role of IGF-1 disruption by PBDE in male reproductive health remains poorly understood. However, the importance of the IGF-1-activated mTOR pathway for testes physiology as well as direct evidence of mTOR modulation by PBDEs suggests the potential involvement of IGF-1/mTOR signaling in PBDE-induced male reproductive toxicity.

5.12. Disruption of Steroidogenesis

Steroidogenesis is the process by which cholesterol is converted into biologically active steroid hormones, including sex steroids [187][188]. The steroidogenic process is regulated mainly by the transcription of genes that encode steroidogenic enzymes and co-factors [189]. For instance, testicular steroidogenic acute regulatory (StAR) protein plays an essential role in the transport of cholesterol to the inner mitochondrial membrane [190], the place where CYP11A1 converts cholesterol to pregnenolone, further used in the endoplasmic reticulum for the production of the final steroid product [191]. Many studies demonstrated that PBDEs and their metabolites affect various steroidogenic enzymes.
The effects of PBDEs on steroidogenic activity were first demonstrated in human adrenocortical carcinoma cells (H295R). An analysis of the expression of ten steroidogenic genes in H295R cells following exposures to DE-71, DE-79, and 20 different PBDE metabolites at 0.025, 0.05, and 0.5 μM concentrations demonstrated the ability of the majority of the tested compounds to change the expression of steroidogenic enzymes with four metabolites (6-MeO-BDE-47, 6-OH-BDE-90, 2′-OH-BDE-68, and 6-OH-BDE-47) producing significant changes at the lowest tested dose [192]. Similarly, most of the PBDEs and their metabolites activated steroidogenic enzymes in another study with the H295R cell line following exposure to hydroxylated and methoxylated PBDEs [193]. Specifically, CYP19 (aromatase) was activated by many metabolites at a 10 μM concentration.
Capitalizing on these studies, the researchers in [194] used a primary culture of rat Leydig cells to develop a mechanistic hypothesis connecting PBDE exposure with altered testosterone production. They hypothesized that PBDE may interfere with the cAMP pathway, which is known to regulate the rate-limiting enzymes of androgen biosynthesis, StAR, and CYP11A1 [195]. To test this hypothesis, Leydig cells were exposed to DE-71 at doses covering approximately the range between 1 and 30 nM. DE-71 stimulated testosterone secretion at 10 nM and higher concentrations. Increased steroidogenesis was associated with the increased production of cAMP, increased expression of StAR, and increased enzymatic activity of CYP11A1. Similarly, the significant upregulation of steroidogenic genes (SCARB1, StAR, and HSD11B1) was observed in Leydig cells cultured with 100 nM BDE-47 in another study [42]. The co-administration of adenylyl cyclase (AC) inhibitor prevented DE-71-stimulated testosterone secretion, suggesting that PBDE regulates (stimulates) steroidogenesis at some mechanism upstream of the cAMP pathway [194].
The expression of several genes involved in testosterone synthesis (Star, Cyp11a1, Hsd3b1, Cyp17a1, and Hsd17b3) was decreased to varying degrees in prepubertal testes following prenatal exposure to 0.2 mg/kg body weight BDE-99 and higher doses in rat [47] and mouse studies [47]. These changes matched the reduced testosterone levels in circulation. Two studies using high doses of BDE-209 also showed the suppression of steroidogenic enzymes in mouse testes [93][128]. These in vivo results seemingly contradict in vitro data discussed in the previous paragraph. This contradiction may originate from the opposing effects of different doses of PBDE. For example, in a study with another type of steroidogenic cells, bovine luteal cells, BDE-99 stimulated the production of progesterone at 0.1–0.3 μM concentrations, but suppressed it at 1–3 μM concentrations [196]. Given that the threshold between the activating and suppressive concentrations of PBDE could be congener-specific, current knowledge does not enable the prediction of the effects of different PBDE mixes and concentrations on steroidogenesis. However, existing knowledge allows the suggestion that, overall, low doses of PBDE stimulate steroidogenesis and higher doses suppress it.

5.13. Mitochondria Disruption and Cell Apoptosis

Mitochondria play a critical role during the process of spermatogenesis [197]. Apart from their role in energy (ATP) production crucial for secretory activity [198], sperm motility [199], and other physiological aspects of reproductive tissues, mitochondria are the major site for steroid hormone production [200] and they play important roles in cell signaling, cell proliferation, and death [201][202]. Specifically, anti-apoptotic and pro-apoptotic proteins in the mitochondria recruit and activate the caspase cascade to regulate testicular apoptosis during the process of spermatogenesis [203].
Several authors hypothesized that the male reproductive effects of PBDE may be mediated via the disruption of mitochondria, which may lead to apoptosis [44][204]. Most of the studies providing evidence in support of this hypothesis report mitochondria-related outcomes at doses significantly exceeding environmental exposure.
For example, the analysis of mitochondrial outcomes in immortalized mouse spermatocytes exposed to 0.1, 1, 10, and 100 μM of BDE-47 for 48 h showed an altered mitochondrial ultrastructure at the highest dose, the decreased expression of mitochondrial proteins Atp5b and Uqcrc1, and decreased Bcl-2, an anti-apoptotic factor of the mitochondrial apoptotic pathway at the two highest doses and altered MMP and decreased ATP production at the three highest doses [204]. Overall, no mitochondrial disruption was found at the 0.1 μM concentration of BDE-47, suggesting that the doses inducing mitochondrial toxicity exceed the internal doses in the general population by around three orders of magnitude. In cultured rat Leydig cells, 10 μM BDE-99 induced the mitochondrial apoptotic pathway [47]. Similarly, the exposure of rat pheochromocytoma PC12 cells to 10 μM or higher concentrations of BDE-47 resulted in the inhibition of mitochondrial fusion (Mfn1 and Mfn2) and fission proteins (Fis1 and phosphorylated Drp1), decreased production of ATP, dissipation of mitochondrial membrane potential (MMP), mitochondrial fragmentation, and activation of apoptosis, while no mitochondrial disruption was induced by 1 μM of BDE-47 [205]. In the study of HepG2 cells exposed to a range of BDE-47 or BDE-154 doses, BDE-47 induced markers of mitochondria disruption and apoptosis at 1 μM and higher doses, while most of the effects of BDE-154 were observed at higher doses (25 μM) [206]. Both flame retardants failed to produce effects at 0.5 μM doses. The low sensitivity of mitochondria to PBDE exposure was also demonstrated in an in vivo study in which sperm MMP was reduced in mice exposed to 500 and 1500 mg/kg body weight BDE-209, but not at smaller doses (10 and 100 mg/kg body weight) [44].


  1. Saghir, S.A. Polybrominated Biphenyls (PBBs). In Reference Module in Biomedical Sciences; Elsevier: Amsterdam, The Netherlands, 2018; ISBN 978-0-12-801238-3.
  2. Allen, J.G.; Gale, S.; Zoeller, R.T.; Spengler, J.D.; Birnbaum, L.; McNeely, E. PBDE Flame Retardants, Thyroid Disease, and Menopausal Status in U.S. Women. Environ. Health 2016, 15, 60.
  3. Lee, H.J.; Kim, G.B. An Overview of Polybrominated Diphenyl Ethers (PBDEs) in the Marine Environment. Ocean Sci. J. 2015, 50, 119–142.
  4. Siddiqi, M.A.; Laessig, R.H.; Reed, K.D. Polybrominated Diphenyl Ethers (PBDEs): New Pollutants-Old Diseases. Clin. Med. Res. 2003, 1, 281–290.
  5. Wang, Z.; Ma, X.; Lin, Z.; Na, G.; Yao, Z. Congener Specific Distributions of Polybrominated Diphenyl Ethers (PBDEs) in Sediment and Mussel (Mytilus Edulis) of the Bo Sea, China. Chemosphere 2009, 74, 896–901.
  6. Suvorov, A.; Takser, L. Facing the challenge of data transfer from animal models to humans: The case of persistent organohalogens. Environmental Health 2008, 7, 58.
  7. Portet-Koltalo, F.; Guibert, N.; Morin, C.; de Mengin-Fondragon, F.; Frouard, A. Evaluation of Polybrominated Diphenyl Ether (PBDE) Flame Retardants from Various Materials in Professional Seating Furnishing Wastes from French Flows. Waste Manag. 2021, 131, 108–116.
  8. Abbasi, G.; Li, L.; Breivik, K. Global Historical Stocks and Emissions of PBDEs. Environ. Sci. Technol. 2019, 53, 6330–6340.
  9. Sjödin, A.; Mueller, J.F.; Jones, R.; Schütze, A.; Wong, L.-Y.; Caudill, S.P.; Harden, F.A.; Webster, T.F.; Toms, L.-M. Serum Elimination Half-Lives Adjusted for Ongoing Exposure of Tri-to Hexabrominated Diphenyl Ethers: Determined in Persons Moving from North America to Australia. Chemosphere 2020, 248, 125905.
  10. Ether, R.D. An Alternatives Assessment for the Flame Retardant Decabromodiphenyl Ether (DecaBDE). 2014. Available online: (accessed on 15 November 2021).
  11. Leonetti, C.; Butt, C.M.; Hoffman, K.; Miranda, M.L.; Stapleton, H.M. Concentrations of Polybrominated Diphenyl Ethers (PBDEs) and 2,4,6-Tribromophenol in Human Placental Tissues. Environ. Int. 2016, 88, 23–29.
  12. Linares, V.; Bellés, M.; Domingo, J.L. Human Exposure to PBDE and Critical Evaluation of Health Hazards. Arch. Toxicol. 2015, 89, 335–356.
  13. Byrne, S.C.; Miller, P.; Seguinot-Medina, S.; Waghiyi, V.; Buck, C.L.; von Hippel, F.A.; Carpenter, D.O. Associations between Serum Polybrominated Diphenyl Ethers and Thyroid Hormones in a Cross Sectional Study of a Remote Alaska Native Population. Sci. Rep. 2018, 8, 2198.
  14. Moreira Bastos, P.; Eriksson, J.; Vidarson, J.; Bergman, A. Oxidative Transformation of Polybrominated Diphenyl Ether Congeners (PBDEs) and of Hydroxylated PBDEs (OH-PBDEs). Environ. Sci. Pollut. Res. Int. 2008, 15, 606–613.
  15. Lavandier, R.; Quinete, N.; Hauser-Davis, R.A.; Dias, P.S.; Taniguchi, S.; Montone, R.; Moreira, I. Polychlorinated Biphenyls (PCBs) and Polybrominated Diphenyl Ethers (PBDEs) in Three Fish Species from an Estuary in the Southeastern Coast of Brazil. Chemosphere 2013, 90, 2435–2443.
  16. Kim, J.S.; Klösener, J.; Flor, S.; Peters, T.M.; Ludewig, G.; Thorne, P.S.; Robertson, L.W.; Luthe, G. Toxicity Assessment of Air-Delivered Particle-Bound Polybrominated Diphenyl Ethers. Toxicology 2014, 317, 31–39.
  17. Trudel, D.; Scheringer, M.; von Goetz, N.; Hungerbühler, K. Total Consumer Exposure to Polybrominated Diphenyl Ethers in North America and Europe. Environ. Sci. Technol. 2011, 45, 2391–2397.
  18. Geyer, H.J.; Schramm, K.-W.; Darnerud, P.O.; Aune, M.; Feicht, A.; Fried, K.W.; Henkelmann, B.; Lenoir, D.; Schmid, P.; McDonald, T.A. Terminal Elimination Half-Lives of the Brominated Flame Retardants TBBPA, HBCD, and Lower Brominated PBDEs in Humans. Organohalogen Compd. 2004, 66, 6.
  19. Johnson-Restrepo, B.; Kannan, K.; Rapaport, D.P.; Rodan, B.D. Polybrominated Diphenyl Ethers and Polychlorinated Biphenyls in Human Adipose Tissue from New York. Environ. Sci. Technol. 2005, 39, 5177–5182.
  20. Hurley, S.; Goldberg, D.; Nelson, D.O.; Guo, W.; Wang, Y.; Baek, H.-G.; Park, J.-S.; Petreas, M.; Bernstein, L.; Anton-Culver, H.; et al. Temporal Evaluation of Polybrominated Diphenyl Ether (PBDE) Serum Levels in Middle-Aged and Older California Women, 2011–2015. Environ. Sci. Technol. 2017, 51, 4697–4704.
  21. Turyk, M.E.; Persky, V.W.; Imm, P.; Knobeloch, L.; Chatterton, R.; Anderson, H.A. Hormone Disruption by PBDEs in Adult Male Sport Fish Consumers. Environ. Health Perspect. 2008, 116, 1635–1641.
  22. Abdelouahab, N.; Ainmelk, Y.; Takser, L. Polybrominated Diphenyl Ethers and Sperm Quality. Reprod. Toxicol. 2011, 31, 546–550.
  23. Kodavanti, P.R.S.; Coburn, C.G.; Moser, V.C.; MacPhail, R.C.; Fenton, S.E.; Stoker, T.E.; Rayner, J.L.; Kannan, K.; Birnbaum, L.S. Developmental Exposure to a Commercial PBDE Mixture, DE-71: Neurobehavioral, Hormonal, and Reproductive Effects. Toxicol. Sci. 2010, 116, 297–312.
  24. Zhang, T.; Zhou, X.; Xu, A.; Tian, Y.; Wang, Y.; Zhang, Y.; Gu, Q.; Wang, S.; Wang, Z. Toxicity of Polybrominated Diphenyl Ethers (PBDEs) on Rodent Male Reproductive System: A Systematic Review and Meta-Analysis of Randomized Control Studies. Sci. Total Environ. 2020, 720, 137419.
  25. Mumford, S.L.; Kim, S.; Chen, Z.; Gore-Langton, R.E.; Boyd Barr, D.; Buck Louis, G.M. Persistent Organic Pollutants and Semen Quality: The LIFE Study. Chemosphere 2015, 135, 427–435.
  26. Albert, O.; Huang, J.Y.; Aleksa, K.; Hales, B.F.; Goodyer, C.G.; Robaire, B.; Chevrier, J.; Chan, P. Exposure to Polybrominated Diphenyl Ethers and Phthalates in Healthy Men Living in the Greater Montreal Area: A Study of Hormonal Balance and Semen Quality. Environ. Int. 2018, 116, 165–175.
  27. Akutsu, K.; Takatori, S.; Nozawa, S.; Yoshiike, M.; Nakazawa, H.; Hayakawa, K.; Makino, T.; Iwamoto, T. Polybrominated Diphenyl Ethers in Human Serum and Sperm Quality. Bull. Environ. Contam. Toxicol. 2008, 80, 345–350.
  28. Yu, Y.-J.; Lin, B.-G.; Liang, W.-B.; Li, L.-Z.; Hong, Y.; Chen, X.-C.; Xu, X.-Y.; Xiang, M.-D.; Huang, S. Associations between PBDEs Exposure from House Dust and Human Semen Quality at an E-Waste Areas in South China-A Pilot Study. Chemosphere 2018, 198, 266–273.
  29. Toft, G.; Lenters, V.; Vermeulen, R.; Heederik, D.; Thomsen, C.; Becher, G.; Giwercman, A.; Bizzaro, D.; Manicardi, G.C.; Spanò, M.; et al. Exposure to Polybrominated Diphenyl Ethers and Male Reproductive Function in Greenland, Poland and Ukraine. Reprod. Toxicol. 2014, 43, 1–7.
  30. Goodyer, C.G.; Poon, S.; Aleksa, K.; Hou, L.; Atehortua, V.; Carnevale, A.; Koren, G.; Jednak, R.; Emil, S.; Bagli, D.; et al. A Case-Control Study of Maternal Polybrominated Diphenyl Ether (PBDE) Exposure and Cryptorchidism in Canadian Populations. Environ. Health Perspect. 2017, 125, 057004.
  31. Koren, G.; Carnevale, A.; Ling, J.; Ozsarfati, J.; Kapur, B.; Bagli, D. Fetal Exposure to Polybrominated Diphenyl Ethers and the Risk of Hypospadias: Focus on the Congeners Involved. J. Pediatr. Urol. 2019, 15, 405.e1–405.e6.
  32. Poon, S.; Koren, G.; Carnevale, A.; Aleksa, K.; Ling, J.; Ozsarfati, J.; Kapur, B.M.; Bagli, D. Association of In Utero Exposure to Polybrominated Diphenyl Ethers With the Risk of Hypospadias. JAMA Pediatr. 2018, 172, 851–856.
  33. Carmichael, S.L.; Herring, A.H.; Sjödin, A.; Jones, R.; Needham, L.; Ma, C.; Ding, K.; Shaw, G.M. Hypospadias and Halogenated Organic Pollutant Levels in Maternal Mid-Pregnancy Serum Samples. Chemosphere 2010, 80, 641–646.
  34. Li, X.; Gao, H.; Li, P.; Chen, W.; Tang, S.; Liu, L.; Zhou, G.; Xia, T.; Wang, A.; Zhang, S. Impaired Sperm Quantity and Motility in Adult Rats Following Gestational and Lactational Exposure to Environmentally Relevant Levels of PBDE-47: A Potential Role of Thyroid Hormones Disruption. Environ. Pollut. 2021, 268, 115773.
  35. Khalil, A.; Parker, M.; Brown, S.E.; Cevik, S.E.; Guo, L.W.; Jensen, J.; Olmsted, A.; Portman, D.; Wu, H.; Suvorov, A. Perinatal Exposure to 2,2′,4′4′ -Tetrabromodiphenyl Ether Induces Testicular Toxicity in Adult Rats. Toxicology 2017, 389, 21–30.
  36. Suvorov, A.; Shershebnev, A.; Wu, H.; Medvedeva, Y.; Sergeyev, O.; Pilsner, J.R. Perinatal Exposure to Low Dose 2,2′,4,4′-Tetrabromodiphenyl Ether (BDE-47) Alters Sperm DNA Methylation in Adult Rats. Reprod. Toxicol. 2018, 75, 136–143.
  37. Pilsner, J.R.; Shershebnev, A.; Wu, H.; Marcho, C.; Dribnokhodova, O.; Shtratnikova, V.; Sergeyev, O.; Suvorov, A. Aging-Induced Changes in Sperm DNA Methylation Are Modified by Low Dose of Perinatal Flame Retardants. Epigenomics 2021, 13, 285–297.
  38. Suvorov, A.; Pilsner, J.R.; Naumov, V.; Shtratnikova, V.; Zheludkevich, A.; Gerasimov, E.; Logacheva, M.; Sergeyev, O. Aging Induces Profound Changes in SncRNA in Rat Sperm and These Changes Are Modified by Perinatal Exposure to Environmental Flame Retardant. Int. J. Mol. Sci. 2020, 21, E8252.
  39. Ellis-Hutchings, R.G.; Cherr, G.N.; Hanna, L.A.; Keen, C.L. Polybrominated Diphenyl Ether (PBDE)-Induced Alterations in Vitamin A and Thyroid Hormone Concentrations in the Rat during Lactation and Early Postnatal Development. Toxicol. Appl. Pharmacol. 2006, 215, 135–145.
  40. Ramhøj, L.; Mandrup, K.; Hass, U.; Svingen, T.; Axelstad, M. Developmental Exposure to the DE-71 Mixture of Polybrominated Diphenyl Ether (PBDE) Flame Retardants Induce a Complex Pattern of Endocrine Disrupting Effects in Rats. PeerJ 2022, 10, e12738.
  41. Kuriyama, S.N.; Talsness, C.E.; Grote, K.; Chahoud, I. Developmental Exposure to Low Dose PBDE 99: Effects on Male Fertility and Neurobehavior in Rat Offspring. Environ. Health Perspect. 2005, 113, 149–154.
  42. Li, Z.; Li, H.; Li, C.; Yan, H.; Ying, Y.; Li, X.; Zhu, Q.; Ge, R.-S.; Wang, Y. Low Dose of Fire Retardant, 2,2′,4,4′-Tetrabromodiphenyl Ether (BDE47), Stimulates the Proliferation and Differentiation of Progenitor Leydig Cells of Male Rats during Prepuberty. Toxicol. Lett. 2021, 342, 6–19.
  43. Stoker, T.E.; Laws, S.C.; Crofton, K.M.; Hedge, J.M.; Ferrell, J.M.; Cooper, R.L. Assessment of DE-71, a Commercial Polybrominated Diphenyl Ether (PBDE) Mixture, in the EDSP Male and Female Pubertal Protocols. Toxicol. Sci. Off. J. Soc. Toxicol. 2004, 78, 144–155.
  44. Tseng, L.-H.; Lee, C.-W.; Pan, M.-H.; Tsai, S.-S.; Li, M.-H.; Chen, J.-R.; Lay, J.-J.; Hsu, P.-C. Postnatal Exposure of the Male Mouse to 2,2′,3,3′,4,4′,5,5′,6,6′-Decabrominated Diphenyl Ether: Decreased Epididymal Sperm Functions without Alterations in DNA Content and Histology in Testis. Toxicology 2006, 224, 33–43.
  45. Miyaso, H.; Nakamura, N.; Matsuno, Y.; Kawashiro, Y.; Komiyama, M.; Mori, C. Postnatal Exposure to Low-Dose Decabromodiphenyl Ether Adversely Affects Mouse Testes by Increasing Thyrosine Phosphorylation Level of Cortactin. J. Toxicol. Sci. 2012, 37, 987–999.
  46. Miyaso, H.; Nakamura, N.; Naito, M.; Hirai, S.; Matsuno, Y.; Itoh, M.; Mori, C. Early Postnatal Exposure to a Low Dose of Decabromodiphenyl Ether Affects Expression of Androgen and Thyroid Hormone Receptor-Alpha and Its Splicing Variants in Mouse Sertoli Cells. PLoS ONE 2014, 9, e114487.
  47. Zhao, T.; Tang, X.; Li, D.; Zhao, J.; Zhou, R.; Shu, F.; Jia, W.; Fu, W.; Xia, H.; Liu, G. Prenatal Exposure to Environmentally Relevant Levels of PBDE-99 Leads to Testicular Dysgenesis with Steroidogenesis Disorders. J. Hazard. Mater. 2022, 424, 127547.
  48. Wang, Y.; Shi, J.; Li, L.; Liu, D.; Li, L.; Tang, C.; Li, J. Adverse Effects of 2,2′,4,4′-Tetrabromodiphenyl Ether on Semen Quality and Spermatogenesis in Male Mice. Bull. Environ. Contam. Toxicol. 2013, 90, 51–54.
  49. Xu, L.; Gao, S.; Zhao, H.; Wang, L.; Cao, Y.; Xi, J.; Zhang, X.; Dong, X.; Luan, Y. Integrated Proteomic and Metabolomic Analysis of the Testes Characterizes BDE-47-Induced Reproductive Toxicity in Mice. Biomolecules 2021, 11, 821.
  50. Wei, Z.; Xi, J.; Gao, S.; You, X.; Li, N.; Cao, Y.; Wang, L.; Luan, Y.; Dong, X. Metabolomics Coupled with Pathway Analysis Characterizes Metabolic Changes in Response to BDE-3 Induced Reproductive Toxicity in Mice. Sci. Rep. 2018, 8, 5423.
  51. Li, X.; Zhu, Y.; Zhang, C.; Liu, J.; Zhou, G.; Jing, L.; Shi, Z.; Sun, Z.; Zhou, X. BDE-209 Induces Male Reproductive Toxicity via Cell Cycle Arrest and Apoptosis Mediated by DNA Damage Response Signaling Pathways. Environ. Pollut. 2019, 255, 113097.
  52. Li, S.; Che, S.; Chen, S.; Ruan, Z.; Zhang, L. Hesperidin Partly Ameliorates the Decabromodiphenyl Ether-Induced Reproductive Toxicity in Pubertal Mice. Environ. Sci. Pollut. Res. Int. 2022, 1–13.
  53. Zhang, Z.; Zhang, X.; Sun, Z.; Dong, H.; Qiu, L.; Gu, J.; Zhou, J.; Wang, X.; Wang, S.-L. Cytochrome P450 3A1 Mediates 2,2′,4,4′-Tetrabromodiphenyl Ether-Induced Reduction of Spermatogenesis in Adult Rats. PLoS ONE 2013, 8, e66301.
  54. Zhang, Z.; Yu, Y.; Xu, H.; Wang, C.; Ji, M.; Gu, J.; Yang, L.; Zhu, J.; Dong, H.; Wang, S.-L. High-Fat Diet Aggravates 2,2′,4,4′-Tetrabromodiphenyl Ether-Inhibited Testosterone Production via DAX-1 in Leydig Cells in Rats. Toxicol. Appl. Pharmacol. 2017, 323, 1–8.
  55. Van der Ven, L.T.M.; van de Kuil, T.; Leonards, P.E.G.; Slob, W.; Cantón, R.F.; Germer, S.; Visser, T.J.; Litens, S.; Håkansson, H.; Schrenk, D.; et al. A 28-Day Oral Dose Toxicity Study in Wistar Rats Enhanced to Detect Endocrine Effects of Decabromodiphenyl Ether (DecaBDE). Toxicol. Lett. 2008, 179, 6–14.
  56. Stoker, T.E.; Cooper, R.L.; Lambright, C.S.; Wilson, V.S.; Furr, J.; Gray, L.E. In Vivo and in Vitro Anti-Androgenic Effects of DE-71, a Commercial Polybrominated Diphenyl Ether (PBDE) Mixture. Toxicol. Appl. Pharmacol. 2005, 207, 78–88.
  57. Sumner, R.N.; Byers, A.; Zhang, Z.; Agerholm, J.S.; Lindh, L.; England, G.C.W.; Lea, R.G. Environmental Chemicals in Dog Testes Reflect Their Geographical Source and May Be Associated with Altered Pathology. Sci. Rep. 2021, 11, 7361.
  58. Sonne, C.; Leifsson, P.S.; Dietz, R.; Born, E.W.; Letcher, R.J.; Hyldstrup, L.; Riget, F.F.; Kirkegaard, M.; Muir, D.C.G. Xenoendocrine Pollutants May Reduce Size of Sexual Organs in East Greenland Polar Bears (Ursus Maritimus). Environ. Sci. Technol. 2006, 40, 5668–5674.
  59. Pizzino, G.; Irrera, N.; Cucinotta, M.; Pallio, G.; Mannino, F.; Arcoraci, V.; Squadrito, F.; Altavilla, D.; Bitto, A. Oxidative Stress: Harms and Benefits for Human Health. Oxid. Med. Cell. Longev. 2017, 2017, 8416763.
  60. Auten, R.L.; Davis, J.M. Oxygen Toxicity and Reactive Oxygen Species: The Devil Is in the Details. Pediatr. Res. 2009, 66, 121–127.
  61. Li, R.; Jia, Z.; Trush, M.A. Defining ROS in Biology and Medicine. React. Oxyg. Species Apex NC 2016, 1, 9–21.
  62. Allamaneni, S.S.R.; Agarwal, A.; Nallella, K.P.; Sharma, R.K.; Thomas, A.J.; Sikka, S.C. Characterization of Oxidative Stress Status by Evaluation of Reactive Oxygen Species Levels in Whole Semen and Isolated Spermatozoa. Fertil. Steril. 2005, 83, 800–803.
  63. Fingerova, H.; Oborna, I.; Novotny, J.; Svobodova, M.; Brezinova, J.; Radova, L. The Measurement of Reactive Oxygen Species in Human Neat Semen and in Suspended Spermatozoa: A Comparison. Reprod. Biol. Endocrinol. 2009, 7, 118.
  64. Henkel, R.; Kierspel, E.; Stalf, T.; Mehnert, C.; Menkveld, R.; Tinneberg, H.-R.; Schill, W.-B.; Kruger, T.F. Effect of Reactive Oxygen Species Produced by Spermatozoa and Leukocytes on Sperm Functions in Non-Leukocytospermic Patients. Fertil. Steril. 2005, 83, 635–642.
  65. Mupfiga, C.; Fisher, D.; Kruger, T.; Henkel, R. The Relationship between Seminal Leukocytes, Oxidative Status in the Ejaculate, and Apoptotic Markers in Human Spermatozoa. Syst. Biol. Reprod. Med. 2013, 59, 304–311.
  66. Mack, S.R.; Everingham, J.; Zaneveld, L.J. Isolation and Partial Characterization of the Plasma Membrane from Human Spermatozoa. J. Exp. Zool. 1986, 240, 127–136.
  67. Poulos, A.; White, I.G. The phospholipid composition of human spermatozoa and seminal plasma. Reproduction 1973, 35, 265–272.
  68. Sanocka, D.; Kurpisz, M. Reactive Oxygen Species and Sperm Cells. Reprod. Biol. Endocrinol. 2004, 2, 12.
  69. Aitken, R.J.; Clarkson, J.S. Cellular Basis of Defective Sperm Function and Its Association with the Genesis of Reactive Oxygen Species by Human Spermatozoa. J. Reprod. Fertil. 1987, 81, 459–469.
  70. Alvarez, J.G.; Touchstone, J.C.; Blasco, L.; Storey, B.T. Spontaneous Lipid Peroxidation and Production of Hydrogen Peroxide and Superoxide in Human Spermatozoa. Superoxide Dismutase as Major Enzyme Protectant against Oxygen Toxicity. J. Androl. 1987, 8, 338–348.
  71. Alvarez, J.G.; Storey, B.T. Differential Incorporation of Fatty Acids into and Peroxidative Loss of Fatty Acids from Phospholipids of Human Spermatozoa. Mol. Reprod. Dev. 1995, 42, 334–346.
  72. Jones, R.; Mann, T.; Sherins, R. Peroxidative Breakdown of Phospholipids in Human Spermatozoa, Spermicidal Properties of Fatty Acid Peroxides, and Protective Action of Seminal Plasma. Fertil. Steril. 1979, 31, 531–537.
  73. Aitken, R.J.; Curry, B.J. Redox Regulation of Human Sperm Function: From the Physiological Control of Sperm Capacitation to the Etiology of Infertility and DNA Damage in the Germ Line. Antioxid. Redox Signal. 2011, 14, 367–381.
  74. de Lamirande, E.; Gagnon, C. Impact of Reactive Oxygen Species on Spermatozoa: A Balancing Act between Beneficial and Detrimental Effects. Hum. Reprod. 1995, 10 (Suppl. 1), 15–21.
  75. Plante, M.; de Lamirande, E.; Gagnon, C. Reactive Oxygen Species Released by Activated Neutrophils, but Not by Deficient Spermatozoa, Are Sufficient to Affect Normal Sperm Motility. Fertil. Steril. 1994, 62, 387–393.
  76. Shi, T.-Y.; Chen, G.; Huang, X.; Yuan, Y.; Wu, X.; Wu, B.; Li, Z.; Shun, F.; Chen, H.; Shi, H. Effects of Reactive Oxygen Species from Activated Leucocytes on Human Sperm Motility, Viability and Morphology. Andrologia 2012, 44, 696–703.
  77. Morielli, T.; O’Flaherty, C. Oxidative Stress Impairs Function and Increases Redox Protein Modifications in Human Spermatozoa. Reprod. Camb. Engl. 2015, 149, 113–123.
  78. Ichikawa, T.; Oeda, T.; Ohmori, H.; Schill, W.B. Reactive Oxygen Species Influence the Acrosome Reaction but Not Acrosin Activity in Human Spermatozoa. Int. J. Androl. 1999, 22, 37–42.
  79. Homa, S.T.; Vessey, W.; Perez-Miranda, A.; Riyait, T.; Agarwal, A. Reactive Oxygen Species (ROS) in Human Semen: Determination of a Reference Range. J. Assist. Reprod. Genet. 2015, 32, 757–764.
  80. Agarwal, A.; Virk, G.; Ong, C.; du Plessis, S.S. Effect of Oxidative Stress on Male Reproduction. World J. Mens Health 2014, 32, 1–17.
  81. Asadi, N.; Bahmani, M.; Kheradmand, A.; Rafieian-Kopaei, M. The Impact of Oxidative Stress on Testicular Function and the Role of Antioxidants in Improving It: A Review. J. Clin. Diagn. Res. JCDR 2017, 11, IE01–IE05.
  82. Bansal, A.K.; Bilaspuri, G.S. Impacts of Oxidative Stress and Antioxidants on Semen Functions. Vet. Med. Int. 2010, 2011, e686137.
  83. Palani, A.; Alahmar, A. Impact of Oxidative Stress on Semen Parameters in Normozoospermic Infertile Men: A Case–Control Study. Afr. J. Urol. 2020, 26, 50.
  84. Deavall, D.G.; Martin, E.A.; Horner, J.M.; Roberts, R. Drug-Induced Oxidative Stress and Toxicity. J. Toxicol. 2012, 2012, e645460.
  85. Lakey, P.S.J.; Berkemeier, T.; Tong, H.; Arangio, A.M.; Lucas, K.; Pöschl, U.; Shiraiwa, M. Chemical Exposure-Response Relationship between Air Pollutants and Reactive Oxygen Species in the Human Respiratory Tract. Sci. Rep. 2016, 6, 32916.
  86. Manuguerra, S.; Espinosa Ruiz, C.; Santulli, A.; Messina, C.M. Sub-Lethal Doses of Polybrominated Diphenyl Ethers, in Vitro, Promote Oxidative Stress and Modulate Molecular Markers Related to Cell Cycle, Antioxidant Balance and Cellular Energy Management. Int. J. Environ. Res. Public. Health 2019, 16, 588.
  87. Zhong, Y.F.; Wang, L.L.; Yin, L.L.; An, J.; Hou, M.L.; Zheng, K.W.; Zhang, X.Y.; Wu, M.H.; Yu, Z.Q.; Sheng, G.Y.; et al. Cytotoxic Effects and Oxidative Stress Response of Six PBDE Metabolites on Human L02 Cells. J. Environ. Sci. Health Part A 2011, 46, 1320–1327.
  88. Zhu, Y.; Jing, L.; Li, X.; Zheng, D.; Zhou, G.; Zhang, Y.; Sang, Y.; Shi, Z.; Sun, Z.; Zhou, X. Decabromodiphenyl Ether Disturbs Hepatic Glycolipid Metabolism by Regulating the PI3K/AKT/GLUT4 and MTOR/PPARγ/RXRα Pathway in Mice and L02 Cells. Sci. Total Environ. 2021, 763, 142936.
  89. Alonso, V.; Linares, V.; Bellés, M.; Albina, M.L.; Pujol, A.; Domingo, J.L.; Sánchez, D.J. Effects of BDE-99 on Hormone Homeostasis and Biochemical Parameters in Adult Male Rats. Food Chem. Toxicol. 2010, 48, 2206–2211.
  90. Zhai, J.-X.; Wang, X.-H.; Zhang, Z.-X.; Zou, L.-W.; Ding, S.-S. Studying the lipid peroxidation index, morphology and apoptosis in testis of male BALB/c mice exposed to polybrominated diphenyl ether (BDE-209). Zhonghua Lao Dong Wei Sheng Zhi Ye Bing Za Zhi Zhonghua Laodong Weisheng Zhiyebing Zazhi Chin. J. Ind. Hyg. Occup. Dis. 2011, 29, 294–298.
  91. Tseng, L.-H.; Hsu, P.-C.; Lee, C.-W.; Tsai, S.-S.; Pan, M.-H.; Li, M.-H. Developmental Exposure to Decabrominated Diphenyl Ether (BDE-209): Effects on Sperm Oxidative Stress and Chromatin DNA Damage in Mouse Offspring. Environ. Toxicol. 2013, 28, 380–389.
  92. Sarkar, D.; Singh, S.K. Maternal Exposure to Polybrominated Diphenyl Ether (BDE-209) during Lactation Affects Germ Cell Survival with Altered Testicular Glucose Homeostasis and Oxidative Status through down-Regulation of Cx43 and P27Kip1 in Prepubertal Mice Offspring. Toxicology 2017, 386, 103–119.
  93. Sarkar, D.; Singh, S.K. Decabromodiphenyl Ether (BDE-209) Exposure to Lactating Mice Perturbs Steroidogenesis and Spermatogenesis in Adult Male Offspring. Ecotoxicol. Environ. Saf. 2021, 209, 111783.
  94. Khalil, A.; Parker, M.; Mpanga, R.; Cevik, S.E.; Thorburn, C.; Suvorov, A. Developmental Exposure to 2,2′,4,4′–Tetrabromodiphenyl Ether Induces Long-Lasting Changes in Liver Metabolism in Male Mice. J. Endocr. Soc. 2017, 1, 323.
  95. Khalil, A.; Cevik, S.E.; Hung, S.; Kolla, S.; Roy, M.A.; Suvorov, A. Developmental Exposure to 2,2′,4,4′-Tetrabromodiphenyl Ether Permanently Alters Blood-Liver Balance of Lipids in Male Mice. Front. Endocrinol. 2018, 9, 548.
  96. Wang, D.; Yan, J.; Teng, M.; Yan, S.; Zhou, Z.; Zhu, W. In Utero and Lactational Exposure to BDE-47 Promotes Obesity Development in Mouse Offspring Fed a High-Fat Diet: Impaired Lipid Metabolism and Intestinal Dysbiosis. Arch. Toxicol. 2018, 92, 1847–1860.
  97. Kozlova, E.V.; Chinthirla, B.D.; Pérez, P.A.; DiPatrizio, N.V.; Argueta, D.A.; Phillips, A.L.; Stapleton, H.M.; González, G.M.; Krum, J.M.; Carrillo, V.; et al. Maternal Transfer of Environmentally Relevant Polybrominated Diphenyl Ethers (PBDEs) Produces a Diabetic Phenotype and Disrupts Glucoregulatory Hormones and Hepatic Endocannabinoids in Adult Mouse Female Offspring. Sci. Rep. 2020, 10, 18102.
  98. Hoppe, A.A.; Carey, G.B. Polybrominated Diphenyl Ethers as Endocrine Disruptors of Adipocyte Metabolism. Obesity 2007, 15, 2942–2950.
  99. Helaleh, M.; Diboun, I.; Al-Tamimi, N.; Al-Sulaiti, H.; Al-Emadi, M.; Madani, A.; Mazloum, N.A.; Latiff, A.; Elrayess, M.A. Association of Polybrominated Diphenyl Ethers in Two Fat Compartments with Increased Risk of Insulin Resistance in Obese Individuals. Chemosphere 2018, 209, 268–276.
  100. Ongono, J.S.; Dow, C.; Gambaretti, J.; Severi, G.; Boutron-Ruault, M.-C.; Bonnet, F.; Fagherazzi, G.; Mancini, F.R. Dietary Exposure to Brominated Flame Retardants and Risk of Type 2 Diabetes in the French E3N Cohort. Environ. Int. 2019, 123, 54–60.
  101. Eslami, B.; Naddafi, K.; Rastkari, N.; Rashidi, B.H.; Djazayeri, A.; Malekafzali, H. Association between Serum Concentrations of Persistent Organic Pollutants and Gestational Diabetes Mellitus in Primiparous Women. Environ. Res. 2016, 151, 706–712.
  102. Boutot, M.E.; Whitcomb, B.W.; Abdelouahab, N.; Baccarelli, A.A.; Boivin, A.; Caku, A.; Gillet, V.; Martinez, G.; Pasquier, J.-C.; Zhu, J.; et al. In Utero Exposure to Persistent Organic Pollutants and Childhood Lipid Levels. Metabolites 2021, 11, 657.
  103. He, Z.; Yin, G.; Li, Q.Q.; Zeng, Q.; Duan, J. Diabetes Mellitus Causes Male Reproductive Dysfunction: A Review of the Evidence and Mechanisms. In Vivo Athens Greece 2021, 35, 2503–2511.
  104. Kahn, B.E.; Brannigan, R.E. Obesity and Male Infertility. Curr. Opin. Urol. 2017, 27, 441–445.
  105. Katib, A. Mechanisms Linking Obesity to Male Infertility. Cent. Eur. J. Urol. 2015, 68, 79–85.
  106. Leisegang, K.; Sengupta, P.; Agarwal, A.; Henkel, R. Obesity and Male Infertility: Mechanisms and Management. Andrologia 2021, 53, e13617.
  107. Zota, A.R.; Geller, R.J.; Romano, L.E.; Coleman-Phox, K.; Adler, N.E.; Parry, E.; Wang, M.; Park, J.-S.; Elmi, A.F.; Laraia, B.A.; et al. Association between Persistent Endocrine-Disrupting Chemicals (PBDEs, OH-PBDEs, PCBs, and PFASs) and Biomarkers of Inflammation and Cellular Aging during Pregnancy and Postpartum. Environ. Int. 2018, 115, 9–20.
  108. Arita, Y.; Yeh, C.; Thoma, T.; Getahun, D.; Menon, R.; Peltier, M.R. Effect of Polybrominated Diphenyl Ether Congeners on Placental Cytokine Production. J. Reprod. Immunol. 2018, 125, 72–79.
  109. Peltier, M.R.; Klimova, N.G.; Arita, Y.; Gurzenda, E.M.; Murthy, A.; Chawala, K.; Lerner, V.; Richardson, J.; Hanna, N. Polybrominated Diphenyl Ethers Enhance the Production of Proinflammatory Cytokines by the Placenta. Placenta 2012, 33, 745–749.
  110. Robinson, J.F.; Kapidzic, M.; Hamilton, E.G.; Chen, H.; Puckett, K.W.; Zhou, Y.; Ona, K.; Parry, E.; Wang, Y.; Park, J.-S.; et al. Genomic Profiling of BDE-47 Effects on Human Placental Cytotrophoblasts. Toxicol. Sci. Off. J. Soc. Toxicol. 2019, 167, 211–226.
  111. Azenabor, A.; Ekun, A.O.; Akinloye, O. Impact of Inflammation on Male Reproductive Tract. J. Reprod. Infertil. 2015, 16, 123–129.
  112. Dutta, S.; Sengupta, P.; Slama, P.; Roychoudhury, S. Oxidative Stress, Testicular Inflammatory Pathways, and Male Reproduction. Int. J. Mol. Sci. 2021, 22, 10043.
  113. Leisegang, K.; Henkel, R. The in Vitro Modulation of Steroidogenesis by Inflammatory Cytokines and Insulin in TM3 Leydig Cells. Reprod. Biol. Endocrinol. 2018, 16, 26.
  114. Zhou, Y.; Zhou, B.; Pache, L.; Chang, M.; Khodabakhshi, A.H.; Tanaseichuk, O.; Benner, C.; Chanda, S.K. Metascape Provides a Biologist-Oriented Resource for the Analysis of Systems-Level Datasets. Nat. Commun. 2019, 10, 1523.
  115. Belloni, V.; Sorci, G.; Paccagnini, E.; Guerreiro, R.; Bellenger, J.; Faivre, B. Disrupting Immune Regulation Incurs Transient Costs in Male Reproductive Function. PLoS ONE 2014, 9, e84606.
  116. Suvorov, A.; Takser, L. Delayed Response in the Rat Frontal Lobe Transcriptome to Perinatal Exposure to the Flame Retardant BDE-47. J. Appl. Toxicol. JAT 2011, 31, 477–483.
  117. Suvorov, A.; Takser, L. Global Gene Expression Analysis in the Livers of Rat Offspring Perinatally Exposed to Low Doses of 2,2′,4,4′-Tetrabromodiphenyl Ether. Environ. Health Perspect. 2010, 118, 97–102.
  118. Fan, L.; Di Ciano-Oliveira, C.; Weed, S.A.; Craig, A.W.B.; Greer, P.A.; Rotstein, O.D.; Kapus, A. Actin Depolymerization-Induced Tyrosine Phosphorylation of Cortactin: The Role of Fer Kinase. Biochem. J. 2004, 380, 581–591.
  119. Anahara, R.; Toyama, Y.; Maekawa, M.; Kai, M.; Ishino, F.; Toshimori, K.; Mori, C. Flutamide Depresses Expression of Cortactin in the Ectoplasmic Specialization between the Sertoli Cells and Spermatids in the Mouse Testis. Food Chem. Toxicol. Int. J. Publ. Br. Ind. Biol. Res. Assoc. 2006, 44, 1050–1056.
  120. Anahara, R.; Toyama, Y.; Mori, C. Flutamide Induces Ultrastructural Changes in Spermatids and the Ectoplasmic Specialization between the Sertoli Cell and Spermatids in Mouse Testes. Reprod. Toxicol. 2004, 18, 589–596.
  121. Berruti, G.; Paiardi, C. The Dynamic of the Apical Ectoplasmic Specialization between Spermatids and Sertoli Cells: The Case of the Small GTPase Rap1. BioMed Res. Int. 2014, 2014, 635979.
  122. Wen, Q.; Tang, E.I.; Li, N.; Mruk, D.D.; Lee, W.M.; Silvestrini, B.; Cheng, C.Y. Regulation of Blood-Testis Barrier (BTB) Dynamics, Role of Actin-, and Microtubule-Based Cytoskeletons. Methods Mol. Biol. Clifton NJ 2018, 1748, 229–243.
  123. Vanholder, R.; Ringoir, S. Artificial Organs—An Overview. Int. J. Artif. Organs 1991, 14, 613–618.
  124. Cheng, C.Y.; Mruk, D.D. The Blood-Testis Barrier and Its Implications for Male Contraception. Pharmacol. Rev. 2012, 64, 16–64.
  125. Gouesse, R.-J.; Lavoie, M.; Dianati, E.; Wade, M.; Hales, B.; Robaire, B.; Plante, I. Gestational and Lactational Exposure to an Environmentally-Relevant Mixture of Brominated Flame Retardants Down-Regulates Junctional Proteins, Thyroid Hormone Receptor A1 Expression and the Proliferation-Apoptosis Balance in Mammary Glands Post Puberty. Toxicol. Sci. 2019, 171, 13–31.
  126. Zhao, Y.; Ao, H.; Chen, L.; Sottas, C.M.; Ge, R.-S.; Zhang, Y. Effect of Brominated Flame Retardant BDE-47 on Androgen Production of Adult Rat Leydig Cells. Toxicol. Lett. 2011, 205, 209–214.
  127. Eskenazi, B.; Rauch, S.A.; Tenerelli, R.; Huen, K.; Holland, N.T.; Lustig, R.H.; Kogut, K.; Bradman, A.; Sjödin, A.; Harley, K.G. In Utero and Childhood DDT, DDE, PBDE and PCBs Exposure and Sex Hormones in Adolescent Boys: The CHAMACOS Study. Int. J. Hyg. Environ. Health 2017, 220, 364–372.
  128. Sarkar, D.; Chowdhury, J.P.; Singh, S.K. Effect of Polybrominated Diphenyl Ether (BDE-209) on Testicular Steroidogenesis and Spermatogenesis through Altered Thyroid Status in Adult Mice. Gen. Comp. Endocrinol. 2016, 239, 50–61.
  129. Johnson, P.I.; Stapleton, H.M.; Mukherjee, B.; Hauser, R.; Meeker, J.D. Associations between Brominated Flame Retardants in House Dust and Hormone Levels in Men. Sci. Total Environ. 2013, 445–446, 177–184.
  130. Hamers, T.; Kamstra, J.H.; Sonneveld, E.; Murk, A.J.; Kester, M.H.A.; Andersson, P.L.; Legler, J.; Brouwer, A. In Vitro Profiling of the Endocrine-Disrupting Potency of Brominated Flame Retardants. Toxicol. Sci. Off. J. Soc. Toxicol. 2006, 92, 157–173.
  131. Kojima, H.; Takeuchi, S.; Uramaru, N.; Sugihara, K.; Yoshida, T.; Kitamura, S. Nuclear Hormone Receptor Activity of Polybrominated Diphenyl Ethers and Their Hydroxylated and Methoxylated Metabolites in Transactivation Assays Using Chinese Hamster Ovary Cells. Environ. Health Perspect. 2009, 117, 1210–1218.
  132. Liu, H.; Hu, W.; Sun, H.; Shen, O.; Wang, X.; Lam, M.H.W.; Giesy, J.P.; Zhang, X.; Yu, H. In Vitro Profiling of Endocrine Disrupting Potency of 2,2′,4,4′-Tetrabromodiphenyl Ether (BDE47) and Related Hydroxylated Analogs (HO-PBDEs). Mar. Pollut. Bull. 2011, 63, 287–296.
  133. Hu, W.; Liu, H.; Sun, H.; Shen, O.; Wang, X.; Lam, M.H.W.; Giesy, J.P.; Zhang, X.; Yu, H. Endocrine Effects of Methoxylated Brominated Diphenyl Ethers in Three in Vitro Models. Mar. Pollut. Bull. 2011, 62, 2356–2361.
  134. Sheikh, I.A. Endocrine-Disrupting Potential of Polybrominated Diphenyl Ethers (PBDEs) on Androgen Receptor Signaling: A Structural Insight. Struct. Chem. 2021, 32, 887–897.
  135. Costa, L.G.; Giordano, G.; Tagliaferri, S.; Caglieri, A.; Mutti, A. Polybrominated Diphenyl Ether (PBDE) Flame Retardants: Environmental Contamination, Human Body Burden and Potential Adverse Health Effects. Acta Bio-Med. Atenei Parm. 2008, 79, 172–183.
  136. Jiang, Y.; Yuan, L.; Lin, Q.; Ma, S.; Yu, Y. Polybrominated Diphenyl Ethers in the Environment and Human External and Internal Exposure in China: A Review. Sci. Total Environ. 2019, 696, 133902.
  137. Meijer, L.; Martijn, A.; Melessen, J.; Brouwer, A.; Weiss, J.; de Jong, F.H.; Sauer, P.J.J. Influence of Prenatal Organohalogen Levels on Infant Male Sexual Development: Sex Hormone Levels, Testes Volume and Penile Length. Hum. Reprod. Oxf. Engl. 2012, 27, 867–872.
  138. Meerts, I.A.; Letcher, R.J.; Hoving, S.; Marsh, G.; Bergman, A.; Lemmen, J.G.; van der Burg, B.; Brouwer, A. In Vitro Estrogenicity of Polybrominated Diphenyl Ethers, Hydroxylated PDBEs, and Polybrominated Bisphenol A Compounds. Environ. Health Perspect. 2001, 109, 399–407.
  139. Kitamura, S.; Shinohara, S.; Iwase, E.; Sugihara, K.; Uramaru, N.; Shigematsu, H.; Fujimoto, N.; Ohta, S. Affinity for Thyroid Hormone and Estrogen Receptors of Hydroxylated Polybrominated Diphenyl Ethers. J. Health Sci. 2008, 54, 607–614.
  140. Mercado-Feliciano, M.; Bigsby, R.M. Hydroxylated Metabolites of the Polybrominated Diphenyl Ether Mixture DE-71 Are Weak Estrogen Receptor-Alpha Ligands. Environ. Health Perspect. 2008, 116, 1315–1321.
  141. Li, X.; Gao, Y.; Guo, L.-H.; Jiang, G. Structure-Dependent Activities of Hydroxylated Polybrominated Diphenyl Ethers on Human Estrogen Receptor. Toxicology 2013, 309, 15–22.
  142. Hamers, T.; Kamstra, J.H.; Sonneveld, E.; Murk, A.J.; Visser, T.J.; Van Velzen, M.J.M.; Brouwer, A.; Bergman, A. Biotransformation of Brominated Flame Retardants into Potentially Endocrine-Disrupting Metabolites, with Special Attention to 2,2′,4,4′-Tetrabromodiphenyl Ether (BDE-47). Mol. Nutr. Food Res. 2008, 52, 284–298.
  143. Shanle, E.K.; Xu, W. Endocrine Disrupting Chemicals Targeting Estrogen Receptor Signaling: Identification and Mechanisms of Action. Chem. Res. Toxicol. 2011, 24, 6–19.
  144. Cao, L.-Y.; Ren, X.-M.; Yang, Y.; Wan, B.; Guo, L.-H.; Chen, D.; Fan, Y. Hydroxylated Polybrominated Biphenyl Ethers Exert Estrogenic Effects via Non-Genomic G Protein–Coupled Estrogen Receptor Mediated Pathways. Environ. Health Perspect. 2018, 126, 057005.
  145. Kester, M.H.A.; Bulduk, S.; van Toor, H.; Tibboel, D.; Meinl, W.; Glatt, H.; Falany, C.N.; Coughtrie, M.W.H.; Schuur, A.G.; Brouwer, A.; et al. Potent Inhibition of Estrogen Sulfotransferase by Hydroxylated Metabolites of Polyhalogenated Aromatic Hydrocarbons Reveals Alternative Mechanism for Estrogenic Activity of Endocrine Disrupters. J. Clin. Endocrinol. Metab. 2002, 87, 1142–1150.
  146. Main, K.M.; Kiviranta, H.; Virtanen, H.E.; Sundqvist, E.; Tuomisto, J.T.; Tuomisto, J.; Vartiainen, T.; Skakkebaek, N.E.; Toppari, J. Flame Retardants in Placenta and Breast Milk and Cryptorchidism in Newborn Boys. Environ. Health Perspect. 2007, 115, 1519–1526.
  147. Meeker, J.D.; Johnson, P.I.; Camann, D.; Hauser, R. Polybrominated Diphenyl Ether (PBDE) Concentrations in House Dust Are Related to Hormone Levels in Men. Sci. Total Environ. 2009, 407, 3425–3429.
  148. Makey, C.M.; McClean, M.D.; Braverman, L.E.; Pearce, E.N.; Sjödin, A.; Weinberg, J.; Webster, T.F. Polybrominated Diphenyl Ether Exposure and Reproductive Hormones in North American Men. Reprod. Toxicol. 2016, 62, 46–52.
  149. Gravel, S.; Lavoué, J.; Bakhiyi, B.; Lavoie, J.; Roberge, B.; Patry, L.; Bouchard, M.F.; Verner, M.-A.; Zayed, J.; Labrèche, F. Multi-Exposures to Suspected Endocrine Disruptors in Electronic Waste Recycling Workers: Associations with Thyroid and Reproductive Hormones. Int. J. Hyg. Environ. Health 2020, 225, 113445.
  150. Yu, Y.; Lin, B.; Chen, X.; Qiao, J.; Li, L.; Liang, Y.; Zhang, G.; Jia, Y.; Zhou, X.; Chen, C.; et al. Polybrominated Diphenyl Ethers in Human Serum, Semen and Indoor Dust: Effects on Hormones Balance and Semen Quality. Sci. Total Environ. 2019, 671, 1017–1025.
  151. Guo, L.-C.; Pan, S.; Yu, S.; Liu, T.; Xiao, J.; Zhu, B.; Qu, Y.; Huang, W.; Li, M.; Li, X.; et al. Human Sex Hormone Disrupting Effects of New Flame Retardants and Their Interactions with Polychlorinated Biphenyls, Polybrominated Diphenyl Ethers, a Case Study in South China. Environ. Sci. Technol. 2018, 52, 13935–13941.
  152. Krassas, G.E.; Poppe, K.; Glinoer, D. Thyroid Function and Human Reproductive Health. Endocr. Rev. 2010, 31, 702–755.
  153. La Vignera, S.; Vita, R.; Condorelli, R.A.; Mongioì, L.M.; Presti, S.; Benvenga, S.; Calogero, A.E. Impact of Thyroid Disease on Testicular Function. Endocrine 2017, 58, 397–407.
  154. Ghiasi, H.; Kaiwar, R.; Nesburn, A.B.; Slanina, S.; Wechsler, S.L. Baculovirus-Expressed Glycoprotein E (GE) of Herpes Simplex Virus Type-1 (HSV-1) Protects Mice against Lethal Intraperitoneal and Lethal Ocular HSV-1 Challenge. Virology 1992, 188, 469–476.
  155. Laslett, A.L.; Li, L.H.; Jester, W.F.; Orth, J.M. Thyroid Hormone Down-Regulates Neural Cell Adhesion Molecule Expression and Affects Attachment of Gonocytes in Sertoli Cell-Gonocyte Cocultures. Endocrinology 2000, 141, 1633–1641.
  156. Mendeluk, G.R.; Rosales, M. Thyroxin Is Useful to Improve Sperm Motility. Int. J. Fertil. Steril. 2016, 10, 208–214.
  157. Suvorov, A.; Girard, S.; Lachapelle, S.; Abdelouahab, N.; Sebire, G.; Takser, L. Perinatal Exposure to Low-Dose BDE-47, an Emergent Environmental Contaminant, Causes Hyperactivity in Rat Offspring. Neonatology 2009, 95, 203–209.
  158. Abdelouahab, N.; Suvorov, A.; Pasquier, J.-C.; Langlois, M.-F.; Praud, J.-P.; Takser, L. Thyroid Disruption by Low-Dose BDE-47 in Prenatally Exposed Lambs. Neonatology 2009, 96, 120–124.
  159. Yang, J.; Ma, Y.; Zhang, X.; Liao, X.; Yang, Y.; Sweetman, A.; Li, H. The Potential Association of Polybrominated Diphenyl Ether Concentrations in Serum to Thyroid Function in Patients with Abnormal Thyroids: A Pilot Study. Ann. Palliat. Med. 2021, 10, 9192–9205.
  160. Zhao, X.; Wang, H.; Li, J.; Shan, Z.; Teng, W.; Teng, X. The Correlation between Polybrominated Diphenyl Ethers (PBDEs) and Thyroid Hormones in the General Population: A Meta-Analysis. PLoS ONE 2015, 10, e0126989.
  161. Cao, J.; Lin, Y.; Guo, L.-H.; Zhang, A.-Q.; Wei, Y.; Yang, Y. Structure-Based Investigation on the Binding Interaction of Hydroxylated Polybrominated Diphenyl Ethers with Thyroxine Transport Proteins. Toxicology 2010, 277, 20–28.
  162. Ren, X.M.; Guo, L.-H. Assessment of the Binding of Hydroxylated Polybrominated Diphenyl Ethers to Thyroid Hormone Transport Proteins Using a Site-Specific Fluorescence Probe. Environ. Sci. Technol. 2012, 46, 4633–4640.
  163. Krainick, J.U.; Thoden, U. Methods of pain modulation by electrical stimulation (author’s transl). Langenbecks Arch. Chir. 1976, 342, 75–81.
  164. Marchesini, G.R.; Meimaridou, A.; Haasnoot, W.; Meulenberg, E.; Albertus, F.; Mizuguchi, M.; Takeuchi, M.; Irth, H.; Murk, A.J. Biosensor Discovery of Thyroxine Transport Disrupting Chemicals. Toxicol. Appl. Pharmacol. 2008, 232, 150–160.
  165. Bansal, R.; Tighe, D.; Danai, A.; Rawn, D.F.K.; Gaertner, D.W.; Arnold, D.L.; Gilbert, M.E.; Zoeller, R.T. Polybrominated Diphenyl Ether (DE-71) Interferes with Thyroid Hormone Action Independent of Effects on Circulating Levels of Thyroid Hormone in Male Rats. Endocrinology 2014, 155, 4104–4112.
  166. Lema, S.C.; Dickey, J.T.; Schultz, I.R.; Swanson, P. Dietary Exposure to 2,2′,4,4′-Tetrabromodiphenyl Ether (PBDE-47) Alters Thyroid Status and Thyroid Hormone-Regulated Gene Transcription in the Pituitary and Brain. Environ. Health Perspect. 2008, 116, 1694–1699.
  167. Roberts, S.C.; Bianco, A.C.; Stapleton, H.M. Disruption of Type 2 Iodothyronine Deiodinase Activity in Cultured Human Glial Cells by Polybrominated Diphenyl Ethers. Chem. Res. Toxicol. 2015, 28, 1265–1274.
  168. Hull, K.L.; Harvey, S. Growth Hormone and Reproduction: A Review of Endocrine and Autocrine/Paracrine Interactions. Int. J. Endocrinol. 2014, 2014, 234014.
  169. Tenuta, M.; Carlomagno, F.; Cangiano, B.; Kanakis, G.; Pozza, C.; Sbardella, E.; Isidori, A.M.; Krausz, C.; Gianfrilli, D. Somatotropic-Testicular Axis: A Crosstalk between GH/IGF-I and Gonadal Hormones during Development, Transition, and Adult Age. Andrology 2021, 9, 168–184.
  170. Lee, H.S.; Park, Y.-S.; Lee, J.S.; Seo, J.T. Serum and Seminal Plasma Insulin-like Growth Factor-1 in Male Infertility. Clin. Exp. Reprod. Med. 2016, 43, 97–101.
  171. Suvorov, A.; Battista, M.-C.; Takser, L. Perinatal Exposure to Low-Dose 2,2′,4,4′-Tetrabromodiphenyl Ether Affects Growth in Rat Offspring: What Is the Role of IGF-1? Toxicology 2009, 260, 126–131.
  172. Xu, X.; Yekeen, T.A.; Xiao, Q.; Wang, Y.; Lu, F.; Huo, X. Placental IGF-1 and IGFBP-3 Expression Correlate with Umbilical Cord Blood PAH and PBDE Levels from Prenatal Exposure to Electronic Waste. Environ. Pollut. 2013, 182, 63–69.
  173. Shy, C.-G.; Huang, H.-L.; Chao, H.-R.; Chang-Chien, G.-P. Cord Blood Levels of Thyroid Hormones and IGF-1 Weakly Correlate with Breast Milk Levels of PBDEs in Taiwan. Int. J. Hyg. Environ. Health 2012, 215, 345–351.
  174. Dibble, C.C.; Manning, B.D. Signal Integration by MTORC1 Coordinates Nutrient Input with Biosynthetic Output. Nat. Cell Biol. 2013, 15, 555–564.
  175. Zoncu, R.; Efeyan, A.; Sabatini, D.M. MTOR: From Growth Signal Integration to Cancer, Diabetes and Ageing. Nat. Rev. Mol. Cell Biol. 2011, 12, 21–35.
  176. Uhlén, M.; Fagerberg, L.; Hallström, B.M.; Lindskog, C.; Oksvold, P.; Mardinoglu, A.; Sivertsson, Å.; Kampf, C.; Sjöstedt, E.; Asplund, A.; et al. Proteomics. Tissue-Based Map of the Human Proteome. Science 2015, 347, 1260419.
  177. Boobes, Y.; Bernieh, B.; Saadi, H.; Raafat Al Hakim, M.; Abouchacra, S. Gonadal Dysfunction and Infertility in Kidney Transplant Patients Receiving Sirolimus. Int. Urol. Nephrol. 2010, 42, 493–498.
  178. Deutsch, M.A.; Kaczmarek, I.; Huber, S.; Schmauss, D.; Beiras-Fernandez, A.; Schmoeckel, M.; Ochsenkuehn, R.; Meiser, B.; Mueller-Hoecker, J.; Reichart, B.; et al. Sirolimus-Associated Infertility: Case Report and Literature Review of Possible Mechanisms. Am. J. Transplant. Off. J. Am. Soc. Transplant. Am. Soc. Transpl. Surg. 2007, 7, 2414–2421.
  179. Huyghe, E.; Matsuda, T.; Thonneau, P. Increasing Incidence of Testicular Cancer Worldwide: A Review. J. Urol. 2003, 170, 5–11.
  180. Johnson, E.M.; Anderson, J.K.; Jacobs, C.; Suh, G.; Humar, A.; Suhr, B.D.; Kerr, S.R.; Matas, A.J. Long-Term Follow-up of Living Kidney Donors: Quality of Life after Donation. Transplantation 1999, 67, 717–721.
  181. Li, N.; Cheng, C.Y. Mammalian Target of Rapamycin Complex (MTOR) Pathway Modulates Blood-Testis Barrier (BTB) Function through F-Actin Organization and Gap Junction. Histol. Histopathol. 2016, 31, 961–968.
  182. Mok, K.W.; Mruk, D.D.; Cheng, C.Y. Regulation of Blood-Testis Barrier (BTB) Dynamics during Spermatogenesis via the “Yin” and “Yang” Effects of Mammalian Target of Rapamycin Complex 1 (MTORC1) and MTORC2. Int. Rev. Cell Mol. Biol. 2013, 301, 291–358.
  183. Mok, K.-W.; Mruk, D.D.; Silvestrini, B.; Cheng, C.Y. RpS6 Regulates Blood-Testis Barrier Dynamics by Affecting F-Actin Organization and Protein Recruitment. Endocrinology 2012, 153, 5036–5048.
  184. Mok, K.-W.; Mruk, D.D.; Lee, W.M.; Cheng, C.Y. Rictor/MTORC2 Regulates Blood-Testis Barrier Dynamics via Its Effects on Gap Junction Communications and Actin Filament Network. FASEB J. Off. Publ. Fed. Am. Soc. Exp. Biol. 2013, 27, 1137–1152.
  185. Conn, C.S.; Qian, S.-B. MTOR Signaling in Protein Homeostasis: Less Is More? Cell Cycle 2011, 10, 1940–1947.
  186. Laplante, M.; Sabatini, D.M. Regulation of MTORC1 and Its Impact on Gene Expression at a Glance. J. Cell Sci. 2013, 126, 1713–1719.
  187. Miller, W.L.; Auchus, R.J. The Molecular Biology, Biochemistry, and Physiology of Human Steroidogenesis and Its Disorders. Endocr. Rev. 2011, 32, 81–151.
  188. Whirledge, S.; Cidlowski, J.A. Chapter 5—Steroid Hormone Action. In Yen and Jaffe’s Reproductive Endocrinology, 8th ed.; Strauss, J.F., Barbieri, R.L., Eds.; Elsevier: Philadelphia, PA, USA, 2019; pp. 115–131.e4. ISBN 978-0-323-47912-7.
  189. Bremer, A.A.; Miller, W.L. Chapter 13—Regulation of Steroidogenesis. In Cellular Endocrinology in Health and Disease; Ulloa-Aguirre, A., Conn, P.M., Eds.; Academic Press: Boston, MA, USA, 2014; pp. 207–227. ISBN 978-0-12-408134-5.
  190. Kumar, V.; Chakraborty, A.; Kural, M.R.; Roy, P. Alteration of Testicular Steroidogenesis and Histopathology of Reproductive System in Male Rats Treated with Triclosan. Reprod. Toxicol. 2009, 27, 177–185.
  191. Rone, M.B.; Fan, J.; Papadopoulos, V. Cholesterol Transport in Steroid Biosynthesis: Role of Protein-Protein Interactions and Implications in Disease States. Biochim. Biophys. Acta 2009, 1791, 646–658.
  192. Song, R.; He, Y.; Murphy, M.B.; Yeung, L.W.Y.; Yu, R.M.K.; Lam, M.H.W.; Lam, P.K.S.; Hecker, M.; Giesy, J.P.; Wu, R.S.S.; et al. Effects of Fifteen PBDE Metabolites, DE71, DE79 and TBBPA on Steroidogenesis in the H295R Cell Line. Chemosphere 2008, 71, 1888–1894.
  193. He, Y.; Murphy, M.B.; Yu, R.M.K.; Lam, M.H.W.; Hecker, M.; Giesy, J.P.; Wu, R.S.S.; Lam, P.K.S. Effects of 20 PBDE Metabolites on Steroidogenesis in the H295R Cell Line. Toxicol. Lett. 2008, 176, 230–238.
  194. Wang, K.-L.; Hsia, S.-M.; Mao, I.-F.; Chen, M.-L.; Wang, S.-W.; Wang, P.S. Effects of Polybrominated Diphenyl Ethers on Steroidogenesis in Rat Leydig Cells. Hum. Reprod. 2011, 26, 2209–2217.
  195. Evaul, K.; Hammes, S.R. Cross-Talk between G Protein-Coupled and Epidermal Growth Factor Receptors Regulates Gonadotropin-Mediated Steroidogenesis in Leydig Cells. J. Biol. Chem. 2008, 283, 27525–27533.
  196. Kabakci, R.; Yigit, A.A. Effects of Bisphenol A, Diethylhexyl Phthalate and Pentabrominated Diphenyl Ether 99 on Steroid Synthesis in Cultured Bovine Luteal Cells. Reprod. Domest. Anim. 2020, 55, 683–690.
  197. Vertika, S.; Singh, K.K.; Rajender, S. Mitochondria, Spermatogenesis, and Male Infertility—An Update. Mitochondrion 2020, 54, 26–40.
  198. Midzak, A.S.; Chen, H.; Aon, M.A.; Papadopoulos, V.; Zirkin, B.R. ATP Synthesis, Mitochondrial Function, and Steroid Biosynthesis in Rodent Primary and Tumor Leydig Cells. Biol. Reprod. 2011, 84, 976–985.
  199. Agnihotri, S.K.; Agrawal, A.K.; Hakim, B.A.; Vishwakarma, A.L.; Narender, T.; Sachan, R.; Sachdev, M. Mitochondrial Membrane Potential (MMP) Regulates Sperm Motility. In Vitro Cell. Dev. Biol. Anim. 2016, 52, 953–960.
  200. Miller, W.L. Steroid Hormone Synthesis in Mitochondria. Mol. Cell. Endocrinol. 2013, 379, 62–73.
  201. Antico Arciuch, V.G.; Elguero, M.E.; Poderoso, J.J.; Carreras, M.C. Mitochondrial Regulation of Cell Cycle and Proliferation. Antioxid. Redox Signal. 2012, 16, 1150–1180.
  202. Tait, S.W.G.; Green, D.R. Mitochondria and Cell Signalling. J. Cell Sci. 2012, 125, 807–815.
  203. Mathur, P.P.; D’Cruz, S.C. The Effect of Environmental Contaminants on Testicular Function. Asian J. Androl. 2011, 13, 585–591.
  204. Huang, S.; Wang, J.; Cui, Y. 2,2′,4,4′-Tetrabromodiphenyl Ether Injures Cell Viability and Mitochondrial Function of Mouse Spermatocytes by Decreasing Mitochondrial Proteins Atp5b and Uqcrc1. Environ. Toxicol. Pharmacol. 2016, 46, 301–310.
  205. Dong, L.; Li, P.; Yang, K.; Liu, L.; Gao, H.; Zhou, G.; Zhao, Q.; Xia, T.; Wang, A.; Zhang, S. Promotion of Mitochondrial Fusion Protects against Developmental PBDE-47 Neurotoxicity by Restoring Mitochondrial Homeostasis and Suppressing Excessive Apoptosis. Theranostics 2020, 10, 1245–1261.
  206. Souza, A.O.; Tasso, M.J.; Oliveira, A.M.C.; Pereira, L.C.; Duarte, F.V.; Oliveira, D.P.; Palmeira, C.M.; Dorta, D.J. Evaluation of Polybrominated Diphenyl Ether Toxicity on HepG2 Cells—Hexabrominated Congener (BDE-154) Is Less Toxic than Tetrabrominated Congener (BDE-47). Basic Clin. Pharmacol. Toxicol. 2016, 119, 485–497.
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