Effects of bisphenol A and its analogs on reproductive health: A mini review.

BPS is structurally similar to BPA and is now used in a variety of common consumer products as a BPA alternative. BPS has been detected in food, indoor dust, personal care products, sediment, and paper products such as currency and cashier’s receipt [ 32 , 154 - 156 ]. BPS exposure frequently occurs through ingestion, inhalation, and dermal contact. BPS was detected in 81% of the human urine samples in the United States and in seven Asian countries with a mean concentration of 0.654 ng/mL, which was comparable to BPA [ 33 ]. Neither the metabolic nor biological fate of BPS has been fully examined. An in vitro study indicated that glucuronidation was the major metabolic pathway for BPS [ 157 ]. As a potential EDC, BPS has been examined in the National Toxicology Program (NTP) Tox21 High Throughput Screening (HTS) Program and was classified as an estrogen agonist with a weak affinity for the ER [ 158 , 159 ]. In addition, Rochester and Bolden reviewed the endocrine activity of BPS by comparing BPS’s estrogenic binding potency against estradiol (EE 2 ), a positive control, on a plethora of estrogenic receptors. They then ran an experiment to examine BPA’s estrogenic potency and compared this value to EE 2 ’s potency. The average estrogenic potency for BPS was calculated to be 0.32 ± 0.28 by dividing the estrogenic potency of BPA from the estrogenic binding potency of BPS, indicating BPS had a similar affinity to ER receptors as BPA [ 160 ]. This study also indicated that BPS was in the same order of magnitude as BPA regarding its androgenic, antiandrogenic, antiestrogenic and aryl hydrocarbon binding affinity, and in inhibitory hormone signaling of adipocytes[ 160 ]. A study conducted by Rosenmai et al. found that BPS bound to estrogen receptors and affected estrogenic and antiandrogenic activity in a similar manner to BPA [ 161 ]. However, there have only been a few in vivo studies that have examined the effects of BPS exposure on female and male reproductive systems (overview in Table 1 ). Yamasaki et al. reported that in the immature rat utero-trophic assay, BPS exposure significantly increased absolute and relative uterine wet weights and blotted weights at doses of 20 and 500 mg/kg, but not at a dose of 200 mg/kg [ 162 ]. In SD adult rats, BPS exposure induced testicular reactive oxygen species (ROS) and lipid peroxidation, altered seminiferous epithelium morphology, and decreased antioxidant enzyme activity and plasma T levels in the highest dose group (50 μg/kg) in spermatogonia, spermatocytes, and spermatids [ 163 ]. However, there were no significant changes in the populations of these testicular cells [ 163 ]. As vertebrates, zebrafish have conserved pharmacological targets and nervous system structures, which is comparable with mammals. This therefore makes zebrafish an ideal specimen to study, as they have a large genetic database, short lifespan, and high fecundity. Zebrafish have been widely used as a model to identify targets as well as modes of the action of EDCs [ 12 ]. In two studies that examined zebrafish, both adult and developmental exposure to BPS resulted in increased plasma estradiol (E2) levels in both sexes, while adult exposure altered the E2 levels at lower doses when compared to developmental exposure. Additionally, male fish exhibited more sensitive responses when compared to the female fish [ 164 , 165 ]. BPS exposure in male zebrafish at different points in their lifespan had decreased T levels [ 164 , 165 ]. Chen et al. reported that Caenorhabditis elegans exposed to BPS showed increased germline apoptosis and activation of DNA damage checkpoint kinase CHK-1. BPS is also known to cause distinct alterations of gene expression at the whole transcriptome level compared to BPA [ 166 ]. Consequentially, these current studies have provided compelling evidence that BPS exposure altered hormone homeostasis and impaired reproductive organs in different sexes, exposure windows, and species. More in vivo studies are needed to validate these current findings and examine the effects of BPS exposure on multiple rodent strains. In an in vitro study, Eladak et al. reported that BPS exposure decreased basal testosterone (T) secretion in mouse or human fetal testis explants, starting at doses of 100 nM and 1000 nM, respectively [ 153 ]. When compared to BPA, BPS treatment induced more significant changes in T secretion in the mouse explants, but fewer changes in human explants [ 153 ]. In an in vitro test on the Leydig cell line, MA-10, a dose of 10 μM BPS significantly increased P4 and P5 levels and increased gene expression levels of CYP51 and 5α-Red1 [ 167 ]. In addition, BPS exposure also decreased cell viability and increased early DNA damage responses and abnormal cytoskeleton structure at a dose of 50 μM in a mouse spermatogonial cell line C18–4 [ 120 ]. The dose responses obtained of cell viability, cell cycle alteration, DNA damage, and cytoskeleton were comparable to BPA, but less drastic when compared to BPAF or TBBPA [ 120 ]. BPF is a component of epoxy resins widely used in tank and pipe linings, industrial floors, coatings, dental sealants, and food packaging materials [ 160 ]. BPF has been found in indoor dust and various food items such as beverages, dairy products, meats, seafood, cereals, fruits, and vegetables [ 32 , 155 ]. It has had been reported that BPF was metabolized and transformed to nonactive sulfates, rather than glucuronides, and excreted through urine in pregnant rats. A study found that active BPF was distributed to multiple tissues, including the liver, uterus, placenta and fetus [ 168 ]. In the United States population, BPF was found in approximately 60% of urine samples, and the concentrations ranged from 0.15 to 0.54 μg/L [ 169 ]. The average estrogenic hormonal potency for BPF, as compared to BPA, was reported to be 1.07 ± 1.20, calculated in the same manner as BPS [ 160 ]. Therefore, BPF may have an effect on the endocrine system equal to or possibly more potent than that of BPA. Limited data are available for BPF exposure and reproductive outcomes in experimental animals (overview in Table 2 ). Two studies conducted on female Wistar rats suggested that short-time postnatal exposure to BPF at doses over 100 mg/kg increased uterine relative weight [ 162 , 170 ]. However, another study did not find any changes in reproductive organ weights in young adult female SD rats exposed to BPF at concentrations reaching 500 mg/kg [ 171 ]. Meanwhile, in male SD rats, BPF exposure increased testis weight without significant spermatological changes at a dose of 500 mg/kg. In the same study, rats in both sexes showed decreased serum triiodothyronine (T3) levels and increased thyroxine (T4) levels. The opposite trends in T3 and T4 levels and no changes to serum thyroid-stimulating hormone (TSH) levels suggested that these observed effects were not related to the endocrine-mediated effects of BPF [ 171 ]. According to the results of recent in vitro studies, BPF’s effects on steroidogenesis are equivocal to BPA. BPF has been reported to increase T secretion and expression of steroidogenic genes (Cyp51 and 5αRed1) at a dose of 100 μM in MA-10 Leydig cells [ 167 ]. Eladak et al. reported that BPF exposure decreased basal T secretion in mouse fetal testis explants starting at a dose of 1000 nM, which followed a similar dose-response curve as BPA. In human fetal testis explants, BPF treatment reduced T levels at a dose of 10 nM with a non-monotonic dose-response curve (10 nM) [ 153 ]. The discrepancy in the current data could be the result of different biotransformations of BPF into its metabolites the in vitro models and the different species and strains tested. Collectively, the data suggest that BPF exposure may alter steroidogenesis both in vitro and in vivo , but its effects on the reproductive organs, oogenesis, spermatogenesis, and embryonic development remain inconclusive. Future studies are necessary to fully validate the current findings in multiple rodent models and determine the consequences of BPF-induced hormone alterations on reproductive functions. BPAF is used widely as a crosslinking agent and a monomer in the plastics industry. There is limited information available on the occurrence of BPAF in consumer products and human urine or blood samples. In the United States, BPAF has been found in various food items with lower detection rates (6% - 30%) compared to BPA [ 155 ]. BPAF was found to be metabolized primarily through glucuronidation and excreted through feces and urine in SD rats [ 172 , 173 ]. In humans, BPAF was detected in urine samples at concentrations ranging from below the detection level to 3.93 μg/L in China and Saudi Arabia [ 174 , 175 ]. It is important to note that BPAF exhibited estrogenic binding potencies greater than those of BPA [ 36 , 37 ]. Additionally, BPAF exhibited a preferential affinity (three times stronger) for ER β compared to ER α [ 176 ]. Relatively few studies have examined the effects of BPAF exposure on steroidogenesis in vivo (overview in Table 3 ). Two studies performed on zebrafish demonstrated that exposure to BPAF during development led to increases in E2 levels in the female fish [ 177 , 178 ]. The gonadal examination carried out on the zebrafish indicated that exposure to 1 mg/L BPAF induced acellular areas in the testis and retarded oocyte development [ 178 ]. In those same studies, BPAF exposure decreased T levels in male fish and also altered testicular morphology [ 177 , 178 ]. In contrast, no decreases in T production were observed in SD rat dams prenatally exposed to BPAF at a dose of 750 mg/kg [ 179 ]. Furthermore, Li et al. reported that BPAF could be transferred via cord blood and lactation, finally accumulating in the offspring’s testes [ 137 ]. Offspring exposed to BPAF prenatally and postnatally showed a significant increase in testicular T levels and alterations of genes involved in cell differentiation and meiosis [ 180 ]. Feng et al. reported that BPAF exposure decreased serum T level and T biosynthesis (200 mg/kg), and increased serum luteinizing hormone (LH) (50 mg/kg) and follicle-stimulating hormone (FSH) (10 mg/kg) levels in adult male SD rats [ 181 ]. Recently, a study revealed that BPAF exposure uniquely impaired pregnancies and sexual development in rats at doses of ~80 and ~280 mg/kg, whereas BPA exposure did not alter these reproductive endpoints at similar concentrations [ 182 ]. These data suggest that BPAF might be a more potent endocrine disruptor than BPA. In in vitro studies, exposure to BPAF inhibited mouse oocyte maturation at concentrations of 50 and 100 μg/mL and induced cell cycle arrest by the activation of spindle assembly checkpoints [ 45 ]. In a recent multi-parametric high-content analysis (HCA) of mouse spermatogonial cells, BPAF exhibited the highest spermatogonial toxicity when compared to TBBPA, BPA, and BPF [ 86 ]. Exposure to 1 μM BPAF altered the nuclear morphology and induced cell cycle arrest, while a concentration of 10 μM caused cytoskeleton perturbation and multinucleation of cells, with DNA damage recognized at a concentration of 25 μM [ 120 ]. This in vitro study indicated that BPAF exposure may target the spermatogonia cells as observed in vivo in a fertility study [ 182 ]. The effects of BPAF on steroidogenesis are still unclear and likely differ depending on different exposure windows. Additionally, further studies should be conducted to examine the effects of BPAF exposure on hormone levels and multiple reproductive endpoints in various experimental strains, species, and exposure windows. TBBPA’s primary use is as a reactive flame retardant in plastics, paper, textiles, and circuit boards. The general population can be exposed to TBBPA through inhalation, dermal contact, and ingestion of fish and shellfish [ 183 ]. TBBPA is absorbed by the gastrointestinal tract, then metabolized to glucuronides and sulfates, which in turn are excreted through feces in animal models [ 184 ]. Several studies have examined TBBPA exposure levels on the general population. Serum TBBPA levels were below the level of detection (0.03 ng/L) in pregnant Canadian women, but detectable in 5% of Inuit adults with concentrations up to 480 ng/L [ 185 ]. In the United States, TBBPA was found in approximately 35% of human breast milk samples at levels between 50–350 pg/kg/day [ 186 ]. However, the current data demonstrated that TBBPA did not interact with ER α/β or AR in a panel of in vitro bioassays [ 187 ]. Van der Ven et al. established a one-generation reproduction study and a subacute toxicity study to examine dose-response relationship of TBBPA and its effects on the endocrine system [ 188 ]. The subacute toxicity test was conducted as a 28-day repeat dose study according to OECD407 guidelines, including doses of 0, 30, 100, 300 mg TBBPA/kg body weight. In the two-generation study, exposure started 10 weeks or 2 weeks before mating and was continued during mating, pregnancy and lactation. No exposure-related histopathological changes were observed in any of the assessed organs in either experiment. There were also no effects on endpoints of reproduction -- i.e., mating success, number of implantation sites, and litter size. The cauda epididymis sperm count and morphology were not affected in either experiment. However, a significant dose-dependent increase of the weight of the testis and pituitary weight were observed in F1 male animals. Decreased circulating thyroxine (T4) levels were observed in both the subacute and one-generational study [ 188 ]. Similarly, another two-generation reproductive study was performed on SD rats exposed to TBBPA at doses of 0, 10, 100 or 1,000 mg/kg/day in corn oil by oral gavage [ 189 ]. The F0 generation was treated during a premating period of 10 weeks and the 2-week mating period. The only significant decrease occurred in serum T4 levels in both the F0 and F1 generations at a dose of 1000 mg/kg [ 189 ]. Cope et al. reported a two-generation reproduction study on reproductive, developmental and neurobehavioral effects of TBBPA at oral doses of 10, 100, and 1000 mg/kg BW/day in SD rats [ 190 ]. These findings are consistent with other multi-generational studies, as there were no marked effects on various reproductive parameters including time to the vaginal opening and anogenital distance, sperm motility, concentration, and morphology in F1 or F2 generations. The decrease in T4 levels in SD rats exposed to a dose of 1000 mg/kg of TBBA was the only change observed [ 190 ]. Zatecka et al. conducted a two-generational study focused on the trans-generational effects of TBBPA in CD1 outbred mice [ 191 ]. Gestational exposure to TBBPA was through drinking water (35 μg/kg) to generate the F1 offspring. Experimental and control animals of the F1 generation were bred in various conditions to evaluate the trans-generational effects on the reproductive system. Significantly reduced testicular weight, increased prostate weight, and increased seminal vesical weights were observed in the F2 generation when both parents (F1) were treated with TBBPA. However, no changes in sperm parameters were observed, and decreased the thickness of the seminiferous epithelium and increased apoptotic cells in the testes by TUNEL staining were observed in both F1 and F2 males [ 191 ]. Utilizing C57BL/6J mice exposed to the same level of TBBPA during the gestation, lactation, pre-pubertal, and pubertal periods up to the age of 70 days, Zatecka, et al. found TBBPA treatment did not induce significant changes in any of the general reproductive system parameters such as the anogenital distance, reproductive organ weight, and sperm count and morphology. However, significant reductions in T and T3 levels, and ratios of protamine 1/protamine 2, and increased total protamine/DNA ratio in sperm and apoptotic spermatozoa were observed [ 192 ]. A study of B6C3F1/N mice exposed to TBBPA for two years (500 mg/kg) showed an increased incidence of uterine epithelial tumors including adenomas, adenocarcinomas, and malignant mixed Mullerian tumors [ 193 ]. In an experiment utilizing zebrafish, TBBPA exposure decreased egg production (0.047 μM) and increased premature egg production (1.5 μM) [ 194 ]. In contrast, there were no changes in the ovary or uterine weight observed in ICR mice exposed to 1.0% TBBPA in their diet [ 195 ]. In two in vitro studies, TBBPA exposure was reported to increase T secretion in Leydig cell line MA-10 at micromolar levels [ 167 , 196 ]. Ogunbayo et al. reported that TBBPA induced Sertoli cell death via disruption of Ca 2+ signaling and through affecting Ca 2+ transport proteins [ 197 ]. In a recent study with the mouse spermatogonia cell line C18–4, TBBPA exposure exhibited a dose-dependent induction of nuclear morphological changes from the high content analysis (HCA), including nuclear area and nuclear shape (LWR and P2A). TBBPA also induced significant increases of LWR at concentrations of 5, 10 and 25 μM and a significant increase of P2A at a concentration of 25 μM for 72 h. TBBPA treatments of 25 μM for 48 h, and of 10, 25 μM for 72 h led to an increase of cells in the G2/M phase accompanied by an increase of apoptotic cells. Additionally, TBBPA exposure (25 μM) induced DNA damage responses and perturbed cytoskeleton structure [ 120 ]. Compared to BPA and BPS, TBBPA exhibited higher spermatogonial toxicity, including dose- and time-dependent alterations impacting nuclear morphology, cell cycle progression, DNA damage responses, and perturbation of the cytoskeleton [ 120 ]. Taken collectively, current studies indicated that the effects of TBBPA on reproduction varied in a limited number of studies. No severe reproductive dysfunctions were consistently observed, even at the higher doses. Some studies of TBBPA revealed alterations in T3 or T4 levels, though those changes did not accompany any histological changes in the thyroid gland or other reproductive organs. In in vitro studies, TBBPA exposure impacted testicular cells at relatively low doses compared to BPA. Future in vivo animal studies are needed to determine TBBPA’s reproductive toxicity and molecular mechanism.

We performed a literature search to identify journal articles related to BPA, BPS, BPF, BPAF, and TBBPA exposure and reproduction. The research included articles published between 2014 and 2017 for BPA and all years up to 2017 for BPA analogs. An electronic search was performed in Pubmed ( http,//www.ncbi.nlm.nih.gov/pubmed ) and Google Scholar ( https://scholar.google.com/ ). Pubmed was selected to identify journal articles as it is considered a main and reliable literature source. Google Scholar was used as a much broader search engine to collect and analyze any literature that may not have been included in Pubmed. Search terms included: {‘Bisphenol A’ OR ‘BPA’ OR ‘bisphenol S’ OR ‘BPS’ OR ‘4,4’-sulfonyldiphenol’ OR ‘bisphenol F’ OR ‘BPF’ OR ‘4,4’-dihydroxydiphenyl-methane’ OR ‘bisphenol AF’ OR ‘BPAF’ OR ‘hexafluorobisphenol A’ OR ‘tetrabromobisphenol A’ OR ‘TBBPA’ OR ‘2,2’,6,6’-Tetrabromo-4,4’-isopropylidenediphenol’} AND { ‘reproductive’ OR ‘oocyte’ OR ‘ovary’ OR ‘uterus’ OR ‘testes’ OR ‘sperm’ OR ‘Leydig cell’ OR ‘Sertoli cell’ OR ‘steroidogenesis’ OR ‘estradiol’ OR ‘testosterone’ OR ‘follicle-stimulating hormone’ OR ‘luteinizing hormone’ OR ‘thyroid’ OR ‘pregnenolone’}.

This review performed a literature search of epidemiological and experimental animal studies, which examined the reproductive toxicities of BPA analogs. While there were several studies that monitored urinary or serum levels of BPA analogs, the association between BPA analog exposure and reproductive dysfunction have not been explored. Emerging evidence from animal models suggests that exposure to these chemicals could adversely affect reproductive functions, including oocyte and sperm quality, steroidogenesis, and ovary and testis functions. These adverse effects might vary based on different testing species, exposure periods, and duration. It is also worth noting that our lab’s previous study using mouse spermatogonial cells combined with high-content analysis (HCA) revealed BPAF exhibited the highest testicular toxicity (20% maximal effect concentration EC20≈ 10 μM), followed by TBBPA, BPA, and BPS [ 120 ]. BPAF induced a dose-dependent increase of multinucleated germ cells (MNGs). Induction of MNGs has been reported following gestational exposure to di-(n-butyl) phthalate (DBP) [ 198 - 201 ]. The accumulation of MNGs in the testis has been associated with the formation of carcinoma in situ (CIS) cells, the known precursor to testicular germ cell cancer (TGCC) in humans [ 202 ]. Additionally, a recently published in vivo study showed consistent data that BPAF exposure uniquely impaired pregnancies and sexual development in rats compared to BPA at similar doses [ 182 ]. BPS uniquely reduced the reproductive lifespan of, and induced distinct transcriptome changes in, C. elegans [ 166 ]. Therefore, given the similarities of their estrogenic potencies to BPA, current data indicates that BPA analogs, especially BPAF and TBBPA, may disrupt reproductive functions in an ER-independent manner. The studies reviewed here provide insufficient toxicological and epidemiological data to characterize and determine the reproductive effects of BPA analogs. In animal models, absorption, distribution, metabolism, and excretion of BPA analogs, as well as their estrogenic potencies, should be determined and evaluated. The effects of developmental and adult exposure should be characterized in multiple species. Currently, the NTP is in the process of carrying out toxicological studies of BPS and BPAF on rodents, and toxicokinetic studies of BPS following oral and intravenous exposure. In in vitro studies, adverse signaling pathways need to be elucidated. As endocrine disruptors, the potential interactions of bisphenols and hormone receptors on different testicular cell lines and their toxic consequences need to be determined, using gain- and loss-of-function approaches. Human studies analyzing the associations between analog exposure levels and hormone levels, oocyte and sperm quality, and pregnancy outcomes should be examined. The single urine spot sample is usually used in BPA studies, but it does not reflect the long-term chemical exposure or determine the real exposure level the progression of a disease’s development. Future epidemiological studies need to employ an advanced exposure assessment to monitor long-term BPA analog exposure on individuals. These multifaceted data will potentially allow for comparing information across BPA analogs and a better risk assessment of bisphenols in general. Given the current data gap, the following types of research studies are needed in the future: Environmental studies are needed to better elucidate the environmental occurrence of BPA analogs and determine sources and pathways of human exposure. Epidemiological studies should be conducted in the general population with precise measurement of chronic exposure level. Co-exposure should also be considered. In vivo studies are needed to better elucidate the metabolic pathways of BPA analogs and target the internal doses and exposure times that are relevant to human exposure. In vitro studies are needed to elucidate the unique mechanisms and the mode of actions of BPA analogs on the reproductive systems, especially the molecular mechanism of the formation of MNGs by BPAF. Environmental studies are needed to better elucidate the environmental occurrence of BPA analogs and determine sources and pathways of human exposure. Epidemiological studies should be conducted in the general population with precise measurement of chronic exposure level. Co-exposure should also be considered. In vivo studies are needed to better elucidate the metabolic pathways of BPA analogs and target the internal doses and exposure times that are relevant to human exposure. In vitro studies are needed to elucidate the unique mechanisms and the mode of actions of BPA analogs on the reproductive systems, especially the molecular mechanism of the formation of MNGs by BPAF.

Bisphenol A (BPA) is a high production volume (HPV) chemical, commonly used in food packaging materials, dental sealants, medical devices and thermal receipts [ 1 ]. Exposure to BPA is ubiquitous via ingestion, inhalation, and dermal contact [ 2 , 3 ]. The Centers for Disease Control and Prevention (CDC) has reported measurable levels of BPA in urine samples in over 90% of the United States population [ 4 , 5 ]. BPA has demonstrated endocrine disrupting effects by interacting with various physiological receptors, such as estrogen receptor α/β (ER α/β), estrogen-related receptor γ, androgen receptor (AR), and thyroid hormone receptor [ 6 - 13 ]. Numerous studies have investigated the reproductive toxicity of BPA, and extensive reviews were conducted to address the strength of the evidence regarding BPA toxicity [ 14 ]. In 2006, an expert panel composed of the National Institute of Environmental Health Sciences (NIEHS), the National Institute of Dental Craniofacial Research (NIDCR), the U.S. Environmental Protection Agency (EPA), and Commonweal, reviewed human exposure to BPA in vivo and in vitro [ 15 ]. The subpanel of experts that focused on in vivo animal studies found contradictory results among the studies. However, they were confident that BPA impacted the male and female reproductive system [ 14 ]. Peretz et al. summarized studies published from 2007 to 2013 to examine the associations between BPA and adverse reproductive outcomes [ 14 ]. Based on the evidence from experimental animals and human exposure from 2007–2013, the authors concluded that BPA at doses below the LOAEL (50 mg/kg/day) impacted female reproduction and had potential adverse effects on the male reproductive system [ 14 ]. Furthermore, recent epidemiological studies have indicated that BPA exposure may potentially be associated with alterations in hormone levels, impairment of ovary and uterine function, and reduction of sperm quality [ 16 - 20 ]. Current data from experimental studies have suggested that BPA exposure adversely affected oocyte quality and maturation, decreased sperm production and quality, damaged testicular cells, perturbed hormone levels, and disrupted ovary function and uterine morphology in animal models [ 21 - 28 ]. Due to widespread exposure and concerns that BPA is a reproductive toxicant, the public drove manufacturers to abandon the use of BPA and introduce analogous chemicals in baby bottles, sippy cups, and infant formula packaging [ 29 , 30 ]. The U.S. Food and Drug Administration then ruled that BPA would no longer be used in the products mentioned above [ 29 , 30 ]. BPA analogs are being used as crosslinking reagents and flame retardants in the plastics industry to produce “BPA-free” products. However, the usage of these chemicals is expected to rise globally despite a lack of production data for these analogs. Recently, the prevalence of BPA analogs in the environment, foods, consumer products, and human urine samples have been reported [ 31 - 35 ]. With high degrees of structural similarities to BPA, these analogs may potentially have a similar endocrine disrupting capacity and the potential to exert adverse effects on the reproductive system. Emerging evidence suggests that BPA analogs interact with various physiological receptors, such as estrogen receptors α and β, androgenic receptors, and aryl hydrocarbons receptors [ 36 , 37 ]. Compared to BPA, little is known about the reproductive toxicity of these analogs. Here, we have reviewed the current literature on BPA, its analogs, and male/female reproduction to summarize the current state of knowledge and the gaps in that knowledge, and to highlight future research directions that could provide valuable information for toxicity evaluation and risk assessment. In addition, the environmental occurrences, human biomonitoring data, toxicokinetics, and the endocrine disrupting capacity of BPA analogs have also been included. This review is structured into 7 topics: (1) BPA and female reproductive health; (2) BPA and male reproductive health; (3) BPS and reproductive health; (4) BPF and reproductive health; (5) BPAF and reproductive health; (6) TBBPA and reproductive health; (7) conclusion and research needs. This review of recent literature focuses on the effects of BPA and its analogs on the male and female reproductive system. We hope the information and conclusions in this review could direct future studies and be useful in risk assessment and in the formation of regulations regarding BPA and its analogs.