Endocrine-Disrupting Chemicals

#### **Chapter 1**

## G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets of Endocrine Disrupting Chemicals

*Valentine Suteau, Patrice Rodien and Mathilde Munier*

### **Abstract**

Endocrine-disrupting chemicals have received significant concern, since they ubiquitously persist in the environment and are able to induce adverse effects on health, and more particularly on reproductive function. Most of the studies focused on nuclear hormone receptors as mediators of sex steroid hormones signaling. However, there are increasing evidences that peptides hormones of the Hypothalamo-Pituitary-Gonadal axis are targets of endocrine-disrupting chemicals (as Gonadotropin-Releasing Hormone, Follicle-Stimulating Hormone, Luteinizing Hormone…). The majority of these hormones act on G protein-coupled membrane receptors. This review summarizes the effects of endocrine-disrupting chemicals on homeostasis of peptides hormone of Hypothalamo-Pituitary-Gonadal axis and on their G protein-coupled membrane receptors signaling revealed by experimental, clinical, and epidemiological studies in human.

**Keywords:** G-protein coupled hormone receptors, hypothalamic-pituitary-gonadal axis, hormones, endocrine-disrupting chemicals

#### **1. Introduction**

Public concern of endocrine-disrupting chemicals (EDCs) has been rising since the 1990s. EDCs are defined as "an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub)populations" [1–3]. EDCs are found in many products comprising plasticizers, personal care products, pesticides… [1]. Humans are constantly exposed to several different EDCs by ingestion, inhalation, and dermal contact. Some classes of EDCs have been studied in detail. Here, we selected three classes of EDCs based on knowledge of their effects on Hypothalamo-Pituitary-Gonadal (HPG) axis: bisphenol A (BPA), phthalates and dichlorodiphenyltrichloroethane (DDT). BPA is one of the most massively produced EDC with over three million tons manufactured annually [4]. It is used in food packaging, toys, resins used in canned, and medical equipment. Because its incomplete

polymerization and its release from polycarbonate at high temperature, exposure to BPA is important *via* food containers [5–7]. Phthalates are used as liquid plasticizers found in a wide range of products including plastics, coatings, toys, cosmetics, and medical tubing. They are classified in two groups: high molecular weight phthalates, such as diethylhexyl phthalte (DEHP), and low molecular weight phthalates, such as dibutyl phthalates (DBP) [8]. DDT, an organochlorine pesticide, was largely used after the Second World War for its insecticidal properties. Although it was banned in the 1970s in the Western World, it continues to be used in developing countries. DDT is a synthetic mixture of three isoforms: p,p'DDT, o,p'DDT and p,p'DDD. EDCs are originally thought to act through nuclear hormone receptors, such as estrogen receptor (ER) or androgen receptor (AR) [9]. During the last decade, we, and others, were interested in the effect of EDCs on G-protein-coupled hormone receptors (GPCRs). These studies have shown that there are chemical compounds in the environment capable of binding to GPCRs and disrupting the activity and intracellular signaling pathways of receptor. Moreover, EDCs may alter pathways involved in hormone biosynthesis and/or receptor signaling regulation. This review summarizes the effects of three classes of EDCs on hormones homeostasis and GPCRs signaling involved in the HPG axis. Several molecular mechanisms can be involved in the EDC effects on the HPG axis. All studies cited here were performed in human species.

#### **2. GPCRs implicated in the HPG axis**

The GPCRs are the largest family of cell-surface receptors with over 800 members accounting for 4% of the encoded human genome [10]. About half of them have sensory functions, mediating olfaction, taste, light perception, and pheromone signaling. The other half (~350–400) are called endo-receptors, i.e. receptors that interact with endogenous ligands [11]. These receptors are involved in the detection of many extracellular stimuli (from photons or ions to large hormones proteins). Thus, they have important roles in various physiological systems. Dysfunction of GPCRs contributes to many human diseases and GPCRs represent 34% of all Food and Drug Administration-approved drugs [12].

GPCRs are characterized by a common structure with seven transmembrane helices with an extracellular N terminus and an intracellular C terminus [13]. The N-terminal portion, or transmembrane domain, constitute the ligand binding site while the C-terminal portion and the intracellular loops form a coupling domain with the intracellular effectors [14].

In the classical GPCR signaling pathway, after ligand binding, activated-GPCR binds the intracellular heterotrimeric G proteins, promoting the release of GDP from the Gα subunit, exchanged for GTP and the dissociation of the GTP-bound α subunit from βγ dimers. The activated G proteins can then transduce and amplify GPCR signals via second messengers to produce a variety of cell responses [15]. Briefly, G*αs* activates adenylyl cyclases to catalyze the conversion of ATP to cAMP. Members of the Gαi family primarily inhibit cAMP production. The Gαq/11 family converts phosphatidylinositol 4,5-bisphosphate to diacylglycerol and inositol 1,4,5-trisphosphate to activate Protein Kinase C and increases intracellular Ca2+ levels. Approximately 10% of GPCRs can be coupled with different types of Gs subunit depending on cell type and context [16]. The second messengers then target other enzymes such as cAMP-dependent protein kinase A (PKA), GMP-dependent protein kinase G (PKG), Ca2+-dependent protein kinase C (PKC) or calciumsensitive enzymes. The Gβγ subunit can also activate a multitude of effectors (GRKs, ion channels, PI3K, phospholipases, MAP kinases) to induce a variety of

#### *G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets… DOI: http://dx.doi.org/10.5772/intechopen.96240*

cellular effects [17]. G protein-mediated signaling is discontinued when the Gα subunit hydrolyzes GTP to GDP, due to its intrinsic GTPase activity. This then leads to the reassociation of Gα with Gβγ to form the inactive heterotrimer [14]. In addition to canonical signaling through heterotrimeric G proteins, some of GPCRs can use alternative modes of GPCR activation and initiate G protein–independent pathway. The main independent pathway involves a coupling with β-arrestin. Originally, β-arrestin was identified as an essential factor in the endocytosis and arrest of GPCR signaling induced by heterotrimeric G proteins. Today, other functions associated with β-arrestins are being studied and coupling to β-arrestins is increasingly described as "scaffolding" proteins involved in multiple G proteinindependent signaling pathways. Indeed, in addition to clathrin, β-arrestins are able to bind to many proteins involved in different signaling pathways (Src, ERK1/2 and JNK3 kinases protein phosphatases, ubiquitin ligases…) [18]. The activation of β-arrestins signaling pathways can take place at the membrane but also in intracellular after internalization [15]. Indeed, a growing amount of evidence suggests that several molecules have not been known to be regulated by G proteins, suggesting that β-arrestin-mediated signaling pathways may be functioning in parallel with G-protein-mediated pathways enhancing GPCR signaling pathways.

The Hypothalamo-Pituitary-Gonadal axis is active in the midgestational fetus and after birth at the minipuberty but is mainly reactivate at onset of puberty. Some receptors of the HPG axis belong to the subfamily of GPCR: gonadotropin-releasing hormone receptor (GnRHR), GPR54/Kisspeptin receptor, Neurokinin B receptor (NK3R), Prokineticin receptor (PROKR2), follicle stimulating hormone receptor (FSHR), human chorionic gonadotropin/luteinizing hormone receptor (hCG/LHR) and Relaxin Family Peptide Receptor 2 (RXFP2).

The GnRH, a neuropeptidic hormone, is secreted by hypothalamic GnRHexpressing neurons into the portal blood vessels in rhythmic pulses [19]. It binds to a membrane receptor, the GnRH receptor, also known as the luteinizing hormone releasing hormone receptor (LHRHR), on pituitary gonadotropic cells and stimulates the biosynthesis and secretion of LH and FSH [19]. GnRHR is predominantly coupled to the Gq-protein [20]. GnRH/GnRHR pathway constitutes the initial step in the HPG axis and controls reproduction in both sexes. GnRH loss-of-function mutations are associated to normosmic hypogonadotropic hypogonadism [21]. GnRH neurons appear to be directly regulated by Kisspeptin-1 (KISS1), with Neurokinin B (NKB) and Prokineticin 2 (PROK2). KISS1 is a peptidic hormone mostly expressed in the hypothalamus [22]. It activates GPR54/KISS1R, which results in the activation of phospholipase C *via* Gq [12]. GPR54 has been described in brain regions, including hypothalamus, but also in peripheral regions [22]. Kisspeptin/GPR54 pathway has a crucial role in the onset of puberty, the regulation of sex hormone mediated secretion of FSH/LH, and in the control of fertility [22, 23]. Inactivating and activating mutations in *KISS1* or *GPR54* genes have been associated with hypogonadotropic hypogonadism and precocious puberty, respectively [23].

Gonadal function is under pituitary control *via* the gonadotropin hormones: follicle stimulating hormone (FSH) and luteinizing hormone (LH) [24]. FSH and LH are synthesized and secreted by the pituitary gonadotropic cells and work together in the reproductive system. The human chorionic gonadotropin (hCG) is secreted by the placenta and controls ovarian function during gestation. LH and hCG share the same GPCR, the hCG/LHR. The FSH and hCG/LH receptor belong to the glycoprotein-hormone receptor family. Activation of the LH and FSH receptor results in the production of intracellular cyclic AMP (cAMP) *via* Gαs proteins [25, 26]. However, FSHR and LHR can also couple to several other effectors such as Gαq and β-arrestin [26–28]. FSHR is expressed in Sertoli and granulosa cells in male and female gonads, respectively, and is required for normal spermatogenesis and growth and maturation of ovarian follicles, as well as for estrogen production [29]. In women, LHR induces luteinization of granulosa cells, progesterone synthesis and *corpus luteum* maintenance during the luteal phase [30]. In men, LH stimulates testosterone production by Leydig cells [30].

Steroid hormones (estrogen, progesterone, and testosterone) secreted by the gonads, bind, and activate nuclear receptors. However, a membrane associated estrogen receptor (GPER) has been identified 15 years ago [31, 32]. Activation of GPER induces intracellular calcium mobilization, cAMP production and phosphorylation cascade involving ERK1/2, PKA, PI3K [33]. This receptor is implicated in many physiological functions: uterine proliferation, metabolism, cardiovascular, immune, and neural system.

More recently, the INSL3/RXFP2 system pathway was identified for its role in reproduction. Insulin-like peptide-3 (INSL3) belongs to the insulin/relaxin family of peptidic hormones [34, 35]. This hormone is mainly produced by testicular Leydig cells and the production is dependent on the state of Leydig cell differentiation [34]. INSL3 is considered as a marker for Leydig cells function. Its best characterized role is in the control of testicular descent since *INSL3* gene inactivation males have bilateral cryptorchidism with testis remaining in abdominal position [36, 37].

#### **3. Effects of EDCs on signaling of HPG axis G-protein coupled receptors**

Effects of EDCs on the activity of HPG axis GPCR identified in the literature search are summarized in **Table 1**.

#### **3.1 Hypothalamic hormones receptor**

Currently, there are no data on the effects of EDCs on the activity of human hypothalamic hormone receptors. However, some studies have been conducted with animal models. Exposure to phthalates leads to a modulation of GnRHR expression (positive or negative depending on the studies) [50, 51], as well as an increase in its expression in rat uterus [52].

#### **3.2 Gonadotropin hormones receptor**

EDCs, like phthalates, increase the FSHR expression in human granulosa cells [38]. DDT has been shown to disturb the FSH induced-cAMP accumulation [39] and aromatase activity in human granulosa cells [40]. Recently, we showed that DDT behaves as an FSHR positive allosteric modulator [41]. DDT interacts with the receptor in the minor binding pocket in the transmembrane domain. DDT acts on the early steps of activation of the FSHR and induces an increase in FSH-stimulated cAMP production. Moreover, the binding of DDT enhances the FSHR response to hCG. The increased response to FSH in the presence of DDT and the gain of sensitivity to hCG may therefore by deleterious. In opposite, BPA is a FSHR negative allosteric modulator [41].

As for FSHR, EDCs, like BPA, disturbes the expression of hCG/LHR in human endometrial stromal cells [42]. In CHO-K1 cells stably transfected with hCG/LHR, DDT reduced the cAMP accumulation induced by hCG [39, 41] and hLH (Munier et al., Arch Toxicol, in revision). Moreover, DDT decreases the hCG- and hLHpromoted β-arrestin 2 recruitment (Munier et al., Arch Toxicol, in revision). DDT seems to act as a negative allosteric modulator of the hCG/LHR signaling.


*G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets… DOI: http://dx.doi.org/10.5772/intechopen.96240*

#### **Table 1.**

*Experimental studies studying the effect of EDC on HPG axis GPCR signaling.*

#### **3.3 Insl3 receptor, RXFP2**

Only one study has very recently focused on the effect if EDCs on receptor signaling to INSL3: RXFP2. In a cellular model of HEK293 transiently expressing human RXFP2, individually, BPA, DEHP and DBP potentiate the cAMP response to INSL3 [43]. Because of their ubiquity, BPA, DEHP and DBP are present in many human biological fluids, as the amniotic liquid. Furthermore, everyone is chronically exposed to mixtures of environmental chemical factors resulting in toxicological interactions that cannot be predicted by reprotoxicological studies of single molecules. The combination of these three molecules, at concentrations found in human amniotic fluid, decreases the basal activity of RXFP2 as well as the response to INSL3. The structural similarity between FSHR and RXFP2 suggests that small hydrophobic molecules, like phthalates and BPA, could use the same binding sites as DDT in FSHR. The binding of one or two compounds to this site could lead to a stabilization of the active state of the receptor driving an increase of agonist activity [53]. In contrast, the binding of three compounds (DEHP+DBP + BPA) likely leads to a steric hindrance that may prevent the conformational changes necessary for the activation of RXFP2 and probably stabilize an inactive state. This study shows that in addition to individual EDC targets, HPG axis GPCRs can also be targeted by EDC cocktails.

#### **3.4 Membrane sexual steroid hormones receptor**

The G protein-coupled receptor (GPER/GPR30) is a membrane estrogen receptor [31]. Gene inactivation of *GPER* in mice did not induce major modifications in reproductive function [54]. However, several studies show that this receptor has pro-oncogenic effects in hormone-dependent cancers. Although many EDCs exhibit low binding affinities to the nuclear ERs and often require relatively high concentrations (>1 μM) to affect genomic pathways, several studies have focused on non-genomic signaling mediated by GPER [55].

Various DDT derivatives and BPA bind to GPER with a Kd between 1 to 10 μM and are competitors of E2 [46]. The binding affinity of EDCs for GPER is higher than for the nuclear receptors. Nevertheless, low concentrations of o,p'DDE and BPA increased cAMP production by GPER [32, 45, 46]. BPA and phthalate (MEHP) also affect proliferation and migration in human cervical cancer cells [56], in human seminoma cells [45], human breast cancer cells and cancer-associated fibroblasts that lack nuclear ERs [47, 57] as well as the migration and invasion of lung cancer cells [48]. BPA modifies these cellular responses by modulating different intracellular signaling pathways (ERK1/2 or Akt phosphorylation, gene expression) through GPER activation. In opposite, GPER mediates BPA-induced intracellular stress generation (ROS production and calcium accumulation) and apoptosis (caspase activation and mitochondrial membrane potential decrease) in human granulosa cells [49]. Recently, it has also been shown that BPA increases GPER gene expression in breast cancer cell lines [44]. Finally, bisphenols AF and B, two substitutes of BPA, exert high estrogenic effects via GPER pathway at nanomolar concentrations [58, 59].

#### **4. Effects of EDCs on the synthesis and secretion of HPG axis hormones**

Effects of EDCs on the synthesis and secretion of HPG axis hormones identified in the literature search are summarized in **Table 2**.

*G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets… DOI: http://dx.doi.org/10.5772/intechopen.96240*




*G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets… DOI: http://dx.doi.org/10.5772/intechopen.96240*


#### **Table 2.**

*Human biomonitoring studies addressing the relationship between EDC and hormones of HPG axis.*

#### **4.1 Hypothalamus level**

#### *4.1.1 Kisspeptin*

No data are available on the impact of DDT on Kisspeptin in epidemiological studies in humans.

Interestingly, a study led on 262 mother–child pairs from China found a positive correlation between cord blood levels of BPA and *KISS1* mRNA expression in placenta tissue [60].

For phthalates, linear regression analysis showed increasing trend for kisspeptin secretion with the concentration of urinary phthalates [61].

#### *4.1.2 GnRH*

No epidemiological or experimental studies are available on the possible link between EDC levels and GnRH concentration in human. This is probably explained by the pulsatile nature of its release and the lack of dosage in clinical practice. However, many effects of EDC on GnRH were observed in rodents [93].

#### **4.2 Pituitary level**

DDT is rapidly metabolized in the body to DDE. Thus, in epidemiological studies, DDE is dosed in the blood more often than DDT. In a cohort of men of reproductive age, statistically significant positive association was found between the serum level of DDE and LH or FSH [63]. However, others studies did not reveal any association between DDT and FSH or LH levels in adult men [62, 64, 65]. In peri and postmenopausal women, inverse correlation was found between serum DDT and LH [66]. Moreover, it has been shown that maternal exposure to DDT or DDE, assayed in prenatal serum, induced a reduction of plasma LH in teenage boys, not found for FSH [67]. A study also showed that the serum levels of LH (basal level and after GnRH stimulation) was significantly higher in girls with detectable serum DDE levels than in girls with undetectable DDE [68]. This difference was not found for FSH [68].

For BPA, studies found that higher urinary BPA concentration was associated with significantly higher concentrations of serum LH in healthy young men, with or without association with FSH [69–71]. However, these results were not confirmed in others cohorts of fertile men [73]. Conversely, another study found a positive correlation between urinary BPA concentration and FSH level, without change in LH level in a cohort of infertile men [72]. In women, no association was found between urinary bisphenol A and LH or FSH levels in premenopausal women [74–76]. Moreover, no

*G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets… DOI: http://dx.doi.org/10.5772/intechopen.96240*

association was found in healthy children for LH and FSH [77, 78]. A modest negative correlation was found between urinary BPA concentration and peak of GnRHstimulated FSH levels in girls with idiopathic central precocious puberty, without difference for LH levels [79].

Maternal phthalates exposure (urinary samples collected during second trimester) was not associated with serum LH level or FSH in offspring during mini-puberty in boys and girls [80]. However, positive correlations were observed between different phthalates and serum LH in prepubescent Korean children (for serum DBP or MEHP) [81], in girls with attention-deficit/hyperactivity disorder (for urinary MEP) [78] and in Chinese population (11–88 years, males and females) (for urinary MEHHP levels) [82]. In the same populations, either negative [82, 83] or no effects [78, 81] were observed on FSH level. In men, urinary phthalate metabolites were positively associated with LH and FSH levels [84] in one study while negative association between urinary phthalates concentrations and levels of FSH was found in American men (for MBzP) [85] and in Danish men (for MEHP or %MiNP) [86] without impact on LH.

Altogether, epidemiological data have linked exposure to EDC and LH and/ or FSH level but evidence were often inconclusive. The inconsistent findings may partly be due to differences in the characteristics and sizes of the cohorts and to the different EDC exposure levels among studies.

#### **4.3 Gonadal level**

#### *4.3.1 Sexual steroid hormones*

Many data are already available on the effect of endocrine disruptors on the secretion of sex steroids. Recent reviews list all available studies for DDT [93], BPA [93, 94] or phthalates [93, 95–97].

#### *4.3.2 INSL3*

No data are available on the impact of DDT on INSL3 in humans epidemiological studies.

Several studies showed that INSL3 was negatively impacted by putative phthalate metabolites. The Diisononyl phthalate (DiNP) metabolite, cx7-MMeHP, and the DEHP metabolite, 5cxMEPP, showed significant negative correlations with INSL3 in amniotic fluid for weeks 11–22 [87]. Moreover, serum levels of INSL3 was negatively associated with urinary concentration of mono-2-ethylhexyl phthalate (MEHP) and MBzP among large cohorts of chinese men of reproductive age [88–90]. In adjusted models, quartiles increases in phthalates metabolites correlated with significant decreases in plasma INSL3 levels [88–90]. It has also been shown that maternal serum MEHP concentration (from 23–35 weeks of gestation) was negatively correlated with INSL3 level in cord blood mainly in boys [91].

There is also an inverse correlation between BPA level and concentration of INSL3 [92]. Indeed, in a population of 180 boys born after 34 weeks of gestation (52 cryptorchid and 128 control), cord blood levels of free BPA correlated negatively with INSL3 [92]. In this study, cord blood INSL3 level was also significantly decreased in the cryptorchid group compared with the control group [92].

*Ex vivo* studies on human testicular explant were performed, to study more precisely the effect of endocrine disruptors on the secretion of INSL3.

No data are available on the impact of DDT on INSL3 in humans experimental studies.

The exposure of fetal testis (8–12 weeks) to BPA at 10−8 M and 10−5 M for 72 h [98], significantly depressed the basal INSL3 production compared with control. This treatment also reduced INSL3 mRNA level by more than 20% [99]. However, BPA did not modify hCG or hLH-stimulated INSL3 production [98]. Conversely, in human adult testes, BPA increased significantly INSL3 production by Leydig cells, at a low doses (10−9 M) [100]. Interestingly, its analogs, Bisphenol B and Bisphenol S also increased INSL3 production at 10−9 and 10−8 M. Moreover, BADGE, another bisphenol, dose dependently increased INSL3 after 48 h of exposure. In contrast, BPE dose dependently inhibited INSL3 levels [100].

For phthalates, di-(2-ethylhexyl) phthalate (DEHP) and mono-(2-ethylhexyl) phthalate (MEHP) exposition on organo-cultured adult human testis did not affect Leydig cell INSL3 concentrations [101].

#### **5. Conclusions**

Most epidemiological and experimental studies focus on the effect of EDCs on the expression and secretion of hormones, as well as on the activity of nuclear steroid receptors. However, a few experimental studies have shown that G proteincoupled membrane receptors of the HPG axis are targets of EDCs as well. It can be pointed out that most of the studies analyzing the effects of EDCs on GPCRs of HPG axis have been performed with cell culture systems. *In vitro* models are valuable tools because they are easily manipulated. But the comparison of the effects of EDCs in wild-type and GPCRs- inactivated animal models could provide additional informations on the mode of action of these compounds.

Mechanisms of GPCR disruption by EDCs include: (1) changes in the expression; (2) interaction with transmembrane domain receptor; (3) modulation of intracellular signaling pathways.

The GPCRs of HPG axis, involved in diverse physiological functions, should be considered as possible contributors of the adverse effects of EDCs on reproduction. How their modulation by EDCs contributes to these deleterious effects should be an important field of investigations in the near future.

#### **Acknowledgements**

V.S was supported by funding from La Société Française d'Endocrinologie. MM was supported by funding from La Société Française d'Endocrinologie et de Diabétologie Pédiatrique and Novo Nordisk.

#### **Conflict of interest**

The authors declare no conflict of interest.

*G-Protein Coupled Hormone Receptors of the Hypothalamic-Pituitary-Gonadal Axis are Targets… DOI: http://dx.doi.org/10.5772/intechopen.96240*

#### **Author details**

Valentine Suteau1,2, Patrice Rodien1,2,3 and Mathilde Munier1,2,3\*

1 UMR CNRS 6015, INSERM 1083, University of Angers, France

2 Department of Endocrinology, University Hospital, Angers, France

3 Reference Center for Rare Diseases of Thyroid and Hormone Receptors, Angers, France

\*Address all correspondence to: mathilde.munier@univ-angers.fr

© 2021 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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#### **Chapter 2**

## Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis

*Priya Gupta, Archisman Mahapatra, Anjali Suman and Rahul Kumar Singh*

#### **Abstract**

The hypothalamic–pituitary-gonadal (HPG) axis plays a crucial and integrative role in the mammalian endocrine regulation to maintain homeostasis. The HPG axis is primarily responsible for governing all the hormonal events related to reproductive activity. Endocrine-disrupting chemicals (EDCs) comprise a diverse group of naturally occurring and synthetic compounds that mimic and interfere with the endogenous chemical hormones. Epidemiological investigations have shown increasing evidence of altered development and detrimental effects on reproductive health during the past 50 years associated with endocrine disruptors affecting the HPG axis. The pleiotropic harmful effects of EDCs act through hormone-dependent downstream signaling pathways responsible for gonad development either through direct interaction with steroid hormone receptor or via epigenetic regulation. Hence, this chapter summarizes the biological plausibility of EDCs exposure and elucidates the mechanism of action underlying EDCs affecting the regulatory circuits of the mammalian HPG axis and reproductive function.

**Keywords:** endocrine disrupting chemicals, hypothalmic-pituitary-gonadal axis, reproduction, epigenetic, polycystic ovarian syndrome

#### **1. Introduction**

In the past decade, endocrine disruptor chemicals (EDCs) in human pathophysiology have gained much more attention due to their ability to affect development and reproduction in humans and wildlife. In accordance with the US Environmental Protection Agency (EPA), EDCs can be defined as 'exogenous agents that disrupt the hormone homeostasis by interfering with the synthesis, secretion, transport, metabolism, receptor binding or elimination of endogenous hormone [1–6]. EDCs comprise mainly heterogeneous compounds, including synthetic (chemical) and natural (plant products such as isoflavones) responsible for disrupting the hormone system. Endocrine disruptors may be found easily in almost every product used in our daily life, including detergents, plastic bottles, food, toys, pesticides, insecticides, and flame retardants. The criteria proposed by the European Commission for being in the process of defining any compound as EDCs should at least exhibit three

characters: (1) an endocrine activity, (2) an adverse or deleterious endocrine mediated effect in the exposed subject or its progeny or subpopulations (3) a plausible cause-effect relationship between the two [7, 8]. Most of the EDCs interfere with the endocrine system by binding the hormonal receptor or regulating genomic expression. Increasing evidence has documented that the highest risk is posed during early and postnatal development while forming organ and neural systems [9]. Sometimes, epigenetic changes (DNA methylation or acetylation) or histone modifications are also involved in endocrine disruption [10].

EDCs are highly heterogeneous and can be classified based on their origin: Industrial and household chemicals (dioxins, phthalates, polychlorinated biphenyls (PCBs), alkylphenols, plasticizers, fire retardants), agricultural (insecticides, pesticides, herbicides, fungicides, phytoestrogens), residential {bisphenol A (BPA), polybrominated biphenyls (PBBs), phthalates} and some pharmaceuticals agents {parabens, diethylstilbestrol (DES)} [3, 5, 11]. Heavy metals such as lead, mercury, cadmium, and arsenic are also included in the EDCs long list (**Figure 1**) [10, 12]. According to the Stockholm Convention (2001), both production and usage of persistent organic pollutants (POPs) was restricted [13]. Guidelines for a list of chemicals were developed to store and eliminate them, which was happened in 2008 and 2014 [14]. Initially, almost twelve POPs were designated as "dirty dozen" due to their severe adverse effects on humans and the ecosystem. Hence, their production and use were banned. The dirty dozen included industrial chemicals, pesticides, and by-products such as aldrin, chlordane, dieldrin, endrin, dichlorodiphenyltrichloroethane (DDT), heptachlor, mirex, toxaphene, PCBs, hexachlorobenzene, polychlorinated dibenzo-p-dioxins, and polychlorinated dibenzofurans [9].

#### **2. Human exposure and route of entry of EDCs**

Endocrine-disrupting chemicals may exhibit various exposure routes to enter the human body. Generally, inhalation (e.g., Plasticizers), dietary intake (e.g., Foods), dermal absorption (e.g., Cosmetics), and embryonic exposure (e.g., Transfer from mother) represent the main exposure pathways (**Figure 2**) [5, 15]. Following any of these pathways, EDCs may enter the food chain and accumulate in different tissues [16]. Mostly EDCs are highly lipophilic in nature and hence accumulate in the adipose tissue having a long half-life. These aforementioned

*Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis DOI: http://dx.doi.org/10.5772/intechopen.96330*

features explain the exact reasons for being the accumulation of EDCs for years in the adipose tissue. Humans and other top predators are at the top of the food chain. Hence, due to bioaccumulation and biomagnification, they may store many EDCs, ultimately leading to various adverse consequences. Even lifelong exposure or fetal or neonatal stage exposure may bring about cumulative or additive or synergic effects. Therefore, the timing of exposure is of utmost importance in evaluating adverse effects on the endocrine system.

Endocrine disruptors such as dioxins, PCBs, perfluorinated compounds, and DDT are commonly found in pesticide-contaminated soil or groundwater or industrial waste quickly enter the human body via oral consumption of food or water. Some of the commonly used EDCs (DDT, vinclozolin, pyrethroids, chlorpyrifos) in households, agriculture, or public disease vector control may come in contact with human skin or through inhalation. In addition to this, cosmetics, personal care products, sunscreens, medications (triclosan, paraben, phthalates) that we apply on our skin are also responsible for their uptake into our body [17]. Professional workers using fungicides, pesticides, and chemicals are most prone and at high risk of EDCs exposure.

#### **3. HPG axis: the central regulator**

The mammalian reproductive cycles are mainly controlled by an intricate play between hypothalamus pituitary and gonads [18]. The hypothalamic Gonadotropin-Releasing Hormone (GnRH) neurons regulate reproductive functions in all vertebrates. The hypothalamic–pituitary-gonadal (HPG) axis plays a crucial role in the normal development of the reproductive system by controlling the ovarian as well as uterine cycles in females and also spermatogenesis in males [18, 19]. In response to GnRH release from the hypothalamus, pituitary gonadotropic cells synthesize and release Luteinizing Hormone (LH) and Follicle-Stimulating Hormone (FSH), which travel via the bloodstream to the target organs. LH and FSH are essential as regulators of both ovarian and testicular function. In males, LH stimulates

**Figure 3.** *Regulation of HPG axis.*

testicular Leydig cells to synthesize and secrete testosterone, which in turn maintains spermatogenesis in Sertoli cells through its paracrine action and exerts sexual and anabolic actions. While FSH acts on Sertoli cells to produce androgen-binding protein, which is critical for spermatogenesis initiation and ultimately augments sperm production [20]. In females, the production of GnRH, LH, and FSH via the HPG axis are similar, but the actions of these hormones are different. LH and FSH exert their function on the ovaries to promote follicular maturation, ovulation, corpus luteum development, and estrogen and progesterone production [21, 22]. These hormones also have a role in regulating the uterine (menstrual) cycle to prepare for ovulation and embryo implantation [23]. These gonadal steroid hormones secreted by ovaries and testis can inhibit GnRH's hypothalamic synthesis via a feedback loop, hence playing a vital role in regulating reproductive function. In most mammals, 17β- Estradiol, testosterone, and progesterone are the primary estrogen, androgen, and progestin, respectively, and each of their receptors are expressed abundantly in the hypothalamus (**Figure 3**) [24].

#### **4. Molecular mechanisms of EDCs**

It is widely known that EDCs molecule, either natural or synthetic, follows the classical mechanism of action to mimic or interfere with the action of an endocrine regulated network of vertebrates. This interference can happen through different

#### *Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis DOI: http://dx.doi.org/10.5772/intechopen.96330*

mechanisms, either through genomic or non-genomic actions. So before understanding the mechanism of endocrine disruption, we must have a clear picture of the endocrine system that relies on the synthesis and release of hormones from various endocrine glands and its transport via the bloodstream to the required distant cells and tissues [25]. This process involves complex interacting signaling pathways and hormone receptors to control normal body function. Estrogen receptors (ERs) regulate the transcription of their target genes via multiple pathways, either directly or indirectly. Any changes in ER signaling may lead to adverse consequences such as hormone-dependent cancer, abnormal fetal growth and development, altered metabolism, and sometimes impaired fertility. Although the effect of EDCs is not limited to only the ligand-dependent ER signaling pathway, it is the best-studied of ER targeted endocrine disruption. ER in response to ligand signals through both genomic as well as a non-genomic pathway. Briefly, ER mediates its signal in the genomic pathway by binding directly to estrogen-responsive elements (ERE) or indirectly through coactivators such as SP-1 or AP-1. Although the best well-studied nuclear receptor cofactors belong to the p160 family of coactivators (e.g., SRC-1, SRC-2, and SRC-3) but the cofactor complex that mediates ER signaling is more complicated.

The non-genomic pathway has a rapid response as within minutes of ligand binding, signal transduction occurs. During estrogen ligand, a G-protein coupled receptor (GPCR-30) activates and mediates the signal independent of ERs and stimulates cAMP production, fluctuates intracellular calcium, or lead to MAPK or PI3K signaling cascades events. BPA and DES are extensively studied EDCs, which also induce rapid estrogen signaling via the non-genomic pathway (**Figure 4**).

**Figure 4.** *Genomic and non-genomic pathway of estrogen signaling.*

Steroid hormones are a group of lipophilic molecules that regulate a wide variety of physiological functions starting from fetal life to adulthood [26]. All these steroid hormones are biosynthesized from cholesterol, including glucocorticoids, mineralocorticoids, progestin, estrogen, and androgen, regulating sexual and reproductive development. Steroidogenic enzymes involved in the steroid hormone biosynthesis pathway are considered as the crucial targets for EDCs action. For example, Researches have demonstrated that exposure of Nonylphenol (NP) concentration 10−4 to 10−1 M at several ages with multiple rat generations (N = 3-4 per group) can decrease P450scc activity in testes [27]. Phthalates exposure, which consists of a particular class of pesticides, inhibits testosterone synthesis in Leydig cells due to direct CYP17 inhibition and hence exerts anti-androgenic effects [28]. Furthermore, BPA of different concentrations (10−8 to 10−4 M) have been shown to inhibit the enzymatic activity of 3β-HSD, CYP17A1, and 17β-HSD3 in a dose-dependent manner in both human and rat testicular microsomes [29]. EDCs exposure also inhibits the activity of 5-α reductase, which is one of the main enzymes involved in dihydrotestosterone production from testosterone and hence in the regulation of masculinization of the external genitalia and the prostrate [30].

Aromatase (cyp19a) is known as the potential target action of EDC, and any modulation in its expression and function alters the estrogen production rate and hence disrupts the estrogen-dependent process [31]. A study suggested that BPA exposure in rat testis R2C Leydig cells stimulate aromatase activity and is correlated with upregulation in COX-2 in R2C cells [32].

In addition to steroidogenic enzymes and hormone metabolism, EDCs are also known to affect hormone-related receptor and their expression. EDCs, due to their similarity in structural features with the endogenous estrogen hormones, bind easily with the estrogen receptors and thus modify estrogen-responsive gene expression. EDCs that display binding with ERs include industrial bisphenolics, pharmaceutical chemicals, phytoestrogens, and organochlorine pesticides. Methoxychlor (MTX) and DDT are examples of organochlorine pesticides that exhibit estrogenic activity through binding with ERα and ERβ ligand-binding domains [33]. Thus, as reported, both pesticides adversely affect the female reproductive trait by impairing normal follicle development and stimulating uterine proliferation [34]. The insecticide endosulfan has a similar estrogenic activity that causes ovarian regression in both in vitro and in vivo studies [35]. Endosulfan competes with estradiol for interacting with the estrogen receptor but with lower affinity and, in turn, affects sex-specific gene expression [36]. Endosulfan can also affect the male reproductive system by decreasing the gene expression of testis-related transcription factors (sox9a and wt1) [37]. BPA possesses a binding affinity for ERs subtypes (ERα and ERβ) and is categorized as a prototypical nonsteroidal ER [38].

It is also important to note that EDCs exhibit multiple hormone-binding actions regardless of their binding to hormonal receptors. For instance, DDT is an agonist for the estrogen receptor, but one of the metabolites of DDT shows anti-androgenic activity [39]. Similarly, BPA shows estrogenic and androgenic activity but exhibits an antagonist nature for thyroid hormone [40, 41].

#### **5. Effect of EDCs on female reproductive system**

The female reproductive system's development is credited to folliculogenesis, where adverse biological effects of EDCs can be observed. The primordial follicles finally become primary, pre-antral, and the antral follicle. Environmental toxicants such as BPA, MTX, and phthalates may interfere at any developmental stage of the

*Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis DOI: http://dx.doi.org/10.5772/intechopen.96330*

aforementioned antral follicle growth and hinder the reproduction, sometimes causing infertility. BPA exposure during in-vitro studies has been shown to inhibit mouse antral follicle development [42]. Follicle growth is dependent on the proliferation of theca and granulosa cells [43].

Estrogens are well-known as the gatekeepers of the female reproductive system. The sudden increase during puberty opens the gate to enter the reproductive life, and at menopause, the gate gets closed due to their decreased level. Thus, the environmental chemicals that behave as agonists or antagonists to estrogen hormone may play a role in precocious puberty, polycystic ovary syndrome, delay in menopause, and premature ovarian failure [44]. Increased growth of the endometrium and breast cancer are some of the unwanted side effects in females that can also occur due to EDCs exposure.

#### **5.1 Puberty and breast development**

The growing evidence of EDCs affecting puberty or early breast development in the female has dramatically increased. It has been observed from the studies that the age of menarche has been decreased from 16-17 years to less than 13 years [5]. The increased estrogen-to-androgen ratio due to overexpression of aromatase activity has been shown to cause early or premature breast development and sometimes gynecomastia in boys [45, 46].

#### **5.2 Breast cancer**

Breast cancer is a multifactorial disease [47] and mainly results from timerelated complex interactions between internal and external factors [48]. Although endogenous estrogen has a role in breast tumor genesis, but estrogen-mimicking exogenous EDCs such as PCBs, BPA, DES, and phthalates have substantial impacts on breast development during the perinatal period and also on carcinogenesis in adults [48]. Studies have reported that women with prenatal exposure to DES, a synthetic estrogen, have an increased risk of breast cancer in their later age (≥40 years) [49]. Perinatal BPA exposure at environmentally relevant concentration alters breast development in both outbred mice and rats. The mode of action of estrogen-mimicking EDCs is two-fold; firstly, their action is on the proliferation of stromal cells and, secondly, concerned with epigenetic mechanisms [50]. EDCs affect the stromal cells by interfering with the estrogen signal pathway. HOXB9 is a homeobox-containing gene that plays a vital role in mammary gland development and is associated with breast cancer. Reports have shown that BPA competes with ER and leads to activation of this gene through histone modification and acetylation [50].

#### **5.3 Uterine disease**

Endometriosis and uterine fibroids are the most common female reproductive disorders, having an estimated combined incidence of up to 70% of women. Due to their cryptic nature, many women with either endometriosis or fibroids may remain asymptomatic or undiagnosed and are more likely caused due to environmental endocrine-active compounds. For instance, non-human primates exposed to environmental contaminant TCDD (dioxin) have a higher endometriosis rate. The onset of fibroids occurs mainly after puberty, and this benign uterine tumor regresses after menopause. Fibroids are found to be more sensitive to the estrogen effect [51]. Hence, due to its dependency on estrogen for its growth, the role of environmental estrogen-mimicking EDCs in fibroid disease should be considered. Several reports have documented that developmental exposure to EDCs such as MTX, PCBs, DDT, DES have been implicated in the development of uterine fibroid disease [52].

#### **5.4 Primary ovarian failure (POF)**

About 1% of the female population under 40 years of age suffers from POF, leading to other comorbidities related to reproductive disorder or early menopause [53]. The possible mechanism of POF development includes premature activation of the follicle, blockage of follicle maturation, and acceleration of apoptosis. Several cases of POF have been reported, and EDCs might have some association with its occurrence. Many EDCs are related to multi-oocyte follicles (MOF) mediated by EDCs-induced ERβ agonist action. Administration of BPA (0.1-1,000 μg/kg) to pregnant mice between the critical periods of differentiation (9th-16th day), ovarian cysts appeared in adulthood, which was significantly more in number in the group that received 1 μg/kg BPA [54]. Paraben is another example of EDC that affects folliculogenesis by stimulating anti-mullerian hormone (AMH) mRNA expression and inhibits the early stage of ovarian follicle in newborn rats [44]. Similarly, MTX inhibits folliculogenesis and increases AMH expression in the pre-antral and early-antral follicles [55]. Recent studies have demonstrated that neonatal exposure of DES (3 μg/kg) induces MOF [56].

#### **5.5 Menstrual irregularity**

Studies have shown that fetal and neonatal exposure to EDCs exposure in humans may interfere with hormonal regulation and result in irregular or long cycles of the menstrual cycle [57]. The irregularities in the menstrual cycle may ultimately reduce fecundity. In-utero exposure to estrogenic compounds such as phytoestrogens or BPA in an animal model such as adult mice increases estrous cycle duration. In comparison, perinatal exposure to BPA results in early suspension and irregular cyclic activity [58] that are likely due to a change in LH secretion's hypothalamic control and ovulation [59].

#### **5.6 Polycystic ovarian syndrome (PCOS)**


This disease is a more prevalent endocrine disorder in women, characterized by anovulation and hyperandrogenism. This syndrome is associated with a higher

**Table 1.**

*A summary of the remarkable studies on effects of EDCs on female reproductive systems.*

#### *Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis DOI: http://dx.doi.org/10.5772/intechopen.96330*

prevalence of obesity, insulin resistance, and other metabolic comorbidities. However, this disease's pathogenesis is still not exact, but evidence shows that genetic and environmental factors such as EDCs may contribute to PCOS's clinical development. BPA, a well-known estrogenic and androgenic endocrine disruptor, acts differently to interfere in reproductive functions. In vitro studies have shown that BPA exposure in rat ovarian thecal interstitial cells increases testosterone synthesis.

In contrast, in male rats, BPA competes with androgens to bind on sex hormone-binding globulin (SHBG), increasing serum-free androgen level. BPA exposure during neonatal conditions could lead to PCOS development. Besides, the estrogenic effect of environmental contaminant BPA enhances the risk of hypertension, type2 diabetes, and dyslipidemia (**Table 1**) [60].

#### **6. Effect of EDCs on male reproductive system**

A large number of studies reported the toxic effects of environmental contaminants on male reproductive health. Numerous EDCs present in our environment may have a causative role in testicular dysgenesis syndrome (TDS) in humans. Impaired spermatogenesis, decrease in semen quality, sperm anomalies, hypospadias, ectopic testes, cryptorchidism, and testicular cancer are important risk factors responsible for the symptoms of developmental disorder, TDS and ultimately causing male infertility. During the normal condition, a functional hormonal feedback loop regulates and ensures proper homeostasis. Sometimes, at the time of sexual development, exposure to certain chemicals may disrupt this tightly-regulated hormonal balance. Even short-term exposures can also have adverse effects and may cause infertility [61].

#### **6.1 Semen and sperm quality**

Semen parameters are used to measure sperm quality and are sometimes considered indicators of compromised male fertility [62]. Recently, the adverse health effects of the male reproductive system, especially due to BPA's estrogenic property, have attracted much more attention. BPA is well-known as a testicular toxicant in animal models as it results in decreased sperm quality and motility, oxidative stress increases, and alters steroidogenesis [42]. Few studies were done in occupationally exposed men, and infertile men have also reported a negative correlation between semen quality and urinary BPA levels. The underlying mechanism may be related to increased oxidative stress and disruption in the steroidogenesis pathway [63].

Functional anomalies of sperm may play a crucial role in male infertility. As reported, in comparison with fertile controls, infertile men have higher seminal reactive oxygen species (ROS) levels; hence studies have suggested ROS might have some role in male infertility. Sperm cells do not contain any cytoplasmic defense enzymes that can serve as ROS scavengers. Therefore, a change in lipid peroxidation (LPO) impairs the plasma membrane fluidity and integrity and ultimately leading to loss of sperm function and movement. Exposure to carbaryl causes low LPO concentrations with an increase in ROS and hence may be associated with altered semen quality, especially sperm motility [64].

Furthermore, BPA was also associated with DNA damage, lower sperm count, abnormal sperm morphology, and motion [65]. Regarding the hormonal level, BPA caused higher FSH, lower inhibin B level, and a lower estradiol-to-testosterone ratio [65]. In a study comprising 375 fertile men, BPA was further associated with decreased free androgen index and higher testosterone level [66].


#### **Table 2.**

*A summary of the remarkable studies on effects of EDCs on male reproductive systems.*

Mixtures of polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs), shortly named 'dioxins', produce adverse health effects reported in several studies. For instance, A study have reported a potential association between chemical dioxins exposure and semen quality. Exposed men have a higher percentage of oligospermia and abnormal sperm morphology [67].

#### **6.2 Testicular cancer, cryptorchidism and hypospadias**

Testicular cancer is the most common disorder found in most young men in many countries. A worldwide trend can be observed towards the increasing number of testicular cancers. Exposure to environmental contaminants such as phthalates causes low semen quality, which ultimately may coincide with the development of cryptorchidism, testicular cancer, and hypospadias. The combination of the three aforementioned disorders of the male reproductive system can be termed as "so-called" testicular TDS. Experimental and epidemiological evidence has demonstrated that TDS is an outcome of interference during embryonal physiological programming and gonadal development in fetal life [68, 69]. EDCs such as phthalates, vinclozolin, polybrominated diphenyl ethers (PBDE), and acetaminophen have shown a significant etiopathogenetic role towards the onset of TDS [10]. Despite all these studies, the exact contribution and mechanism behind pesticide and EDC exposure towards testicular cancer development have not been elucidated to date. However, genetic and environmental factors have been documented for the reduced sperm count and disease formation (**Table 2**) [70].

#### **7. Effect of EDCs on fetal and neonatal stages**

Although adult exposure towards EDCs is an essential factor, the duration of EDCs exposure in fetus and neonate is also of primary concern. It plays a crucial factor in determining its fate. For instance, Due to early EDCs exposure, development is compromised because EDCs at the neonatal stage are extremely sensitive and affect the same brain regions, circuits, hormone-sensitive pathways, and receptors like the endogenous hormones. Such effects have been reported for several EDCs across various species and have profound detrimental consequences in developing organisms compared to adults [71, 72]. The extreme sensitivity of

*Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis DOI: http://dx.doi.org/10.5772/intechopen.96330*

developing fetus and neonate has been described briefly in a chapter titled "The Fragile Fetus" by Howard Bern [73]. All those protective mechanisms present in adults like DNA Repair Mechanism, detoxifying enzymes, liver metabolism, competent immune system, and the blood–brain barrier are not fully developed or functional during the fetus or neonatal stage. Even the metabolic rate of developing organisms is much higher than adults, which may be the cause of increased toxicity due to environmental contaminants exposures. Numerous reports have documented that developmental exposure to EDCs can lead to various adverse effects in adults and sometimes cause tumors in endocrine tissues and adverse reproductive consequences in males and females [74]. Pieces of evidence have stated that exposure to environmental triggers such as EDCs during critical stages in fetal sex differentiation and development in utero disrupts reproductive organ differentiation and sometimes leading to intersex variation (IV) conditions. IV can be defined as a morphological and physiological anomaly where an individual is born with a congenital condition such as ambiguous genitalia/ hermaphrodite or pseudohermaphroditism etc. [75]. The most evident case for endocrine disruption in utero that may lead to the onset of adult disease in the newborn is prenatal exposure to DES. Between 1958 and 1976, doctors prescribed synthetic estrogen to pregnant women to prevent miscarriages and premature delivery. It was almost a 4-6million pregnancies that were treated with DES in the US alone. In 1971, DES was linked with a rare gynecologic neoplasm in female offspring of DES-exposed pregnancies [76]. Subsequent studies have documented the link between maternal treatment with DES and cervicovaginal cancer in DES-exposed daughters, usually in their late teens or early 20s. This was the first evidence of transplacental carcinogenesis in humans [51]. Additionally, the offspring of DES-exposed mothers also had functional and anatomical abnormalities of the uterus and fallopian tubes.

#### **8. Effect of EDCs on epigenetic regulations**

The epigenetic modifications can be defined as "heritable and reversible chemical modifications of chromatin, resulting in an adjustment of its activity without a change in the underlying DNA sequence [77]. Notably, epigenetic effects are mediated by those transcription factors that enhance or repress any specific genes' transcription. The well-studied epigenetic modifications include DNA methylation at the cytosine base, post-translational modification of histone proteins (histone acetylation and deacetylation), and non-coding RNAs. Posttranslational modification of histone protein at specific amino acids, for instance, lysine, may alter chromatin's structure and function. Generally, acetylation of histone at lysine position results in activation of transcription by relaxing the chromatin structure, while methylation of lysine depending on the position may lead to activation or repression of gene expression. However, deacetylation may lead to transcriptional repression or silencing of the genes. Non-coding RNA is the transcript of gene sequences that do not code for proteins but regulate its expression in a cis or trans manner mainly involved in some unique functions such as genomic imprinting, X-chromosome inactivation, and developmental patterning and differentiation [70].

The most commonly studied epigenetic mechanism is the EDCs effect on the enzymes that regulate epigenetic patterning, especially the DNA methyl transferase (DNMTs). The endocrine disruptor, Vincolozolin, an androgen receptor antagonist, induced increased expression of dnmt mRNA expression in an in-vivo model through an AR-mediated pathway [78]. Studies have reported that DES

can activate immediately early genes such as c-myc, c-jun, c-fos, and lactoferrin, which are upregulated during childhood [53]. This activation is possible due to the promoter region's hypomethylation of the lactoferrin gene in the adult uterus [79]; however, no such patterns were observed when an adult was exposed during adulthood at the same interval. Upon prenatal exposure to DES, tet1 mRNA expression was significantly decreased in mouse uterus and the same response found in zebrafish gonads upon BPA exposure [21, 80, 81]. Exposure to EDCs during development could alter the perturbation of the genome's epigenetic patterning and may result in adult-onset disease. The epigenetic changes in the ovary have been reported for the organochlorine pesticide MTX. In a study, MTX exposure from embryonic to postnatal days caused hypermethylation in the ERβ promoter regions by performing DNA methylation analysis using bisulfite sequencing and methylation-specific PCR [82].

#### **9. Conclusion**

Research over the last few years exploited the adverse effects of EDCs, which were lesser-known. Some of them were heavily used in the past decade, and some are just enlisted their names on the list of 'emerging pollutants,' and altogether, they are creating an unhealthy world to live on for us as well as for other animals. This chapter summarizes the harmful effects of EDCs exposure on male and female reproductive physiology and also elucidates the possible mechanism of actions. Though numerous articles and reports regarding these EDCs' toxicological effects are available today, still very little is known about their mechanism of action in our body. The poor understanding of their underlying mechanism in hampering one's endocrine system keeps us at bay to fight against it. Be it your morning toothpaste or your hair nourishing shampoo, and you cannot escape from the exposure of EDCs in your daily life. Even processed foods have EDCs in them. Scientists predict that there are many more of this kind of chemical whose effects are yet to be evaluated. Many countries' government bodies are actively monitoring the situations and taking actions accordingly alongside the scientists who are tirelessly working to find out ways to nullify the adverse effects. We can follow the recent resources available in the public domains and choose a healthy way to live to minimize the EDCs exposure. Extra care should be taken in choosing your packaged food materials, cosmetics, plastic made equipment, cooking utensils, fruits, and vegetables. Finally, emphasis should be on the betterment of regulatory systems for introducing new, untested chemicals alongside the continuous use of chemicals already proved for being EDCs. Utmost care should be taken for pregnant women and infants, who are most vulnerable to EDCs exposure.

#### **Acknowledgements**

All authors thank Banaras Hindu University, India for providing necessary resources and support to write the present chapter with the support from University Grant Commission (UGC)-Junior Research Fellowship to PG, AM, and AS. This work received no external funding from any agency.

#### **Conflict of interest**

The author declares that there is no conflict of interest.

*Effect of Endocrine Disrupting Chemicals on HPG Axis: A Reproductive Endocrine Homeostasis DOI: http://dx.doi.org/10.5772/intechopen.96330*

### **Author contributions**

All authors listed have made a substantial contribution in this chapter. And special thanks to AM for making the required illustrations for this chapter. Thanks to RKS for reviewing the manuscript before the final submission.

#### **Author details**

Priya Gupta, Archisman Mahapatra, Anjali Suman and Rahul Kumar Singh\* Molecular Endocrinology and Toxicology Lab (METLab), Department of Zoology, Institute of Science, Banaras Hindu University, Varanasi, India

\*Address all correspondence to: rks.rna@gmail.com

© 2021 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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#### **Chapter 3**

## Environmental Obesogens and Human Health

*Archisman Mahapatra, Priya Gupta, Anjali Suman and Rahul Kumar Singh*

#### **Abstract**

Obesity is an alarming public health concern that contributes to a substantially increased risk of multiple chronic disorders, including diabetes. As per WHO data, in 2016, almost 39% adult population of the world is overweight, 13% of them were obese. There is prominent evidence on the involvement of environmental endocrine-disrupting chemicals, termed obesogens, in the prevalence of this growing worldwide pandemic, obesity. The exaggerated effect of obesogens on endocrine disruption, lipid metabolism and homeostasis, adipocyte functioning, impaired thermogenesis, inflammation, epigenetics, and overall human health will be covered in this chapter. This chapter will discuss the environmental obesogen hypothesis, the epidemiological and experimental evidence of obesogens, its chemical characteristics, and possible mechanism of actions. It will also focus on some recent indications of obesogens and their correlation in COVID-19 disease pathogenesis. This chapter will try to conclude with strategies for identifying the underlying mechanisms of obesogens within model systems and the human body, including future directions.

**Keywords:** endocrine disrupting chemicals, obesogens, obesity, gut microbiota, peroxisome proliferator-activated receptor γ, lipid metabolism, energy homeostasis

#### **1. Introduction**

#### **1.1 Endocrine-disrupting chemicals**

As defined by the US Environmental Protection Agency (EPA) [1], an endocrine-disrupting chemical (EDCs) is "an exogenous agent that interferes with synthesis, secretion, transport, metabolism, binding action, or elimination of natural blood-borne hormones that are present in the body and are responsible for homeostasis, reproduction, and developmental process." Diamanti-Kandarakis et al. [2] among thousands of human-made chemicals, almost 1000 chemicals may have endocrine-disrupting properties [3]. Initially, it was thought that EDCs deploy their actions mainly through various nuclear hormone receptors like estrogen receptors (ERs), progesterone receptors (PRs), androgen receptors (ARs) and thyroid receptors. However, as research progressed on EDCs and their mechanism of actions, it is now known that they can also act on non-nuclear receptors, nonsteroid receptors, orphan receptors and other enzymatic pathways related to metabolism, cancer and other physiological processes [2].

As the compounds classified under EDCs are from dispersed heterogeneous sources, they can be divided into two major classes, synthetic and natural. Synthetic EDCs include industrial solvents and their byproducts [dioxins, polychlorinated biphenyls (PCBs), polybrominated biphenyls (PBBs), alkylphenols etc.], agricultural pesticides [methoxychlor (MTX), chlorpyrifos, dichlorodiphenyltrichloroethane (DDT)], fungicides, herbicides, insecticides, plasticizers [phthalates, bisphenol A (BPA)] and pharmaceuticals [diethylstilbestrol (DES)] whereas, phytoestrogens (genistein, coumestrol etc.) are grouped under natural sources of EDCs. Humans are exposed to the broad range of EDCs mainly through the dietary intake (fish, meat, dairy and poultry products) and to some extent by inhalation and dermal uptake [2, 4]. Mostly EDCs are highly lipophilic, and they tend to get accumulated in the adipose tissues [5, 6]. They can accumulate in human and other large mammals' fatty tissues through biomagnification and bioaccumulation as they are the top predators in the food chain [7]. Due to their long half-life, they remain stored in the adipose tissues for years. Persistent Organic Pollutants (POPs) are the best example of long term accumulations in human tissues [8]. However, plasticizers like BPA have a very short estimated half-life of about four hours. Instead of bioaccumulation, they generally get excreted via urine [9]. Still, BPA has a very adverse effect on the human endocrine system due to their continuous exposure throughout the days [10].

Among the vast range of chemicals under EDCs, some are referred to as "obesogens" as they promote or induce weight gain in individuals by altering endocrine pathways involved in metabolism, energy homeostasis and appetite. The phthalates, perfluorinated compounds, BPA, dioxins, and some pesticides showed obesogenic potentials [11, 12]. Though their mechanism of action is not very well understood, some report indicated that these chemicals might act through Peroxisome proliferator-activated receptor gamma γ (PPAR-γ), a ligandactivated transcription factor, has a role in various cellular functions as well as glucose homeostasis, lipid metabolism, and prevention of oxidative stress [13, 14]. Some suggest they may act via the thyroid axis, as the thyroid hormone is a crucial regulator of metabolism [15, 16]. Hence, this field is relatively new and emerging in EDC's research and needs further studies.

#### **2. Environmental obesogen hypothesis**

The prevalence of obesity and associated diseases like type 2 diabetes, cardiovascular diseases, metabolic syndromes and cancers are progressively increasing at an alarming rate in recent years. Globally the cases of obesity have nearly tripled since 1975. As per a WHO report, in 2016, 13% of adults aged 18 or more are obese worldwide. A more recent report stated that approximately thirty-eight million children (under five years) are obese. In simple language, obesity can be defined as an "abnormal or excessive fat accumulation that may impair health" [17]. The measure of obesity is generally done by body mass index (BMI), defined as a person's weight in kilograms divided by the square of his/her height in meters (kg/m2 ). A BMI of 30 or greater falls within the obese range; the limit changes to 25 or more in Asian populations [18, 19]. It is a widely accepted fact that the primary cause of obesity is the imbalance between calory intake and energy expenditure. However, obesity is a complex disease caused mainly by endocrine disruption, which also involves interaction between genetic and environmental factors.

The Obesogen Hypothesis suggests that environmental chemicals, characterized as "obesogens," induce obesity by enhancing the engagement, differentiation and size of adipocytes, by altering metabolic setpoints or modifying the hormonal control of appetite and satiety [20]. Many EDCs are obesogens in nature and found abundantly in our environment, which may induce adipogenesis and lipid accumulation in the tissues. About 50 of such compounds have been identified to date [20]. Various mechanisms of action of the obesogens are discussed later in this chapter.

### **3. Chemical characteristics of obesogens**

Obesogens have peculiar characteristics which make them potential to interfere with various endocrine and metabolic pathways. They are believed to be xenohormones as they imitate or partially resemble natural hormones and have unwanted physiological effects. They can bind to endocrine receptors present on the cell membrane, cytosol, or nucleus, thereby altering their natural functions [21]. Along with the structural similarities with native hormones, their ability to do this also relies on its lipophilicity and small molecular weight. Partition coefficient, half-life and molecular weight are the three main components of xenohormones. A partition coefficient (P) is "the ratio of the concentration of a substance in one medium or phase (C1) to the concentration in a second phase (C2) when the two concentrations are at equilibrium; that is, partition coefficient = (C1/C2)equal." [22]. This is how the distribution efficiency of a chemical is measured between two mediums. Here in obesogen's case, it is between the tissue and blood. A compound's octanol– water partition coefficient expresses that (KOW), referred to the ratio of a chemical's concentration in the octanol phase to its concentration in the aqueous phase of a two-phase octanol/water system [23]. It is an essential measure of its lipophilicity of a chemical. The bioaccumulation and toxicity of a chemical largely depend upon


#### **Table 1.**

*Chemical characteristics of some obesogens.*

KOW. As being organic, obesogens are naturally lipophilic compound, which means they have a higher KOW value. More the value of the KOW of a compound, the more will be its tendency to accumulate in the adipose tissues [24].

Now, coming to the half-life, the biological half-life of a chemical is the time it takes to break down or eliminate half of the chemical's quantity from the body. In the body, a longer biological half-life implies longer endurance. Ideally, obesogens have longer biological half-lives means a short exposure can have life-long consequences [25]. The last of the three properties, molecular weight, refers to the size of a compound molecule. Small molecules can diffuse more readily through adipocytes. However, many large molecules having high molecular weight can give rise to smaller metabolites which may have a similar effect to obesogens [24]. The bioaccumulation and the binding affinity for the receptors largely depend upon these three criteria. Many obesogens perfectly fit into these criteria. Moreover, some of them are also resistant to degradation [e.g. 2,2′,4,4′-tetrabromodiphenyl ether (BDE-47)] [21]. A summary of some well-known obesogens with their characteristics is listed in **Table 1**.

#### **4. Mechanisms of action of obesogens**

Though the mechanism of obesogens' actions in inducing obesity is not very clear, some studies suggest few mechanisms by which obesogen could act. This disruption of lipid homeostasis by obesogen may involve several mechanisms, some of which are as follows (**Figure 1**):

**Figure 1.** *Mechanisms of obesogen actions.*

*Environmental Obesogens and Human Health DOI: http://dx.doi.org/10.5772/intechopen.96730*


#### **4.1 PPAR**γ**-RXR mediated**

Obesogens generally disturb the endocrine system by interfering with PPARγ and other hormone receptors like estrogen receptor, androgen receptors and glucocorticoid receptors. PPARγ is one of the primary regulators of adipogenesis. It is highly expressed in adipose tissues and induce differentiation of adipocytes by promoting lipogenic enzymes. Along with adipogenesis, it activates genes involved in maintaining energy balance. Upon activation, PPARγ forms a heterodimer complex with nuclear receptor 9-cis retinoic acid receptor (RXR) an act as promoters for the genes required for storage of fatty acid and repression of lipolysis. That is why this PPARγ:RXR heterodimer is called the "master regulator of adipogenesis" [39]. Obesogen tributyltin (TBT) acts as a ligand and show high binding affinity with PPARγ and nuclear receptor RXR. By activating PPARγ and RXR, it might promote adipogenesis and lipid dysbiosis [40, 41]. Obesogens like spirodiclofen and quinoxyfen activate PPARγ while others like fludioxonil activate RXR [13]. Phthalates are also known activators of PPARγ, as they are shown to promote 3 T3-L1 cells to adipocytes differentiation [42]. Obesogen can increase the amount of adipose tissue by increasing the size as well as numbers of adipocytes. They can induce the Mesenchymal Stem Cells (MSCs) to differentiate into preadipocytes and adipocytes [43]. In vitro assays show numerous compounds with obesogenic properties can induce the Mesenchymal Stem Cells (MSCs) to differentiate into preadipocytes and adipocytes via PPARγ dependent pathways. TBT exposure to 3 T3-L1 preadipocytes induces them to differentiate into white adipose tissues (WAT) [44]. Bisphenol A (BPA), combined with insulin, can accelerate the conversion of 3 T3-L1 fibroblasts to adipocytes [45]. Even prenatal exposure to TBT in mouse shows preferential differentiation of MSCs towards the adipose lineage [43] (**Figure 2**).

From the studies available so far, it is evident that any ligand which can bind to PPARγ can induce adipogenesis and can be called obesogens. However, as human adipose tissue stores many of them, they can have a more significant cumulative effect. These additive effects are not well studied yet.

#### **4.2 Other receptor-mediated**

Obesogens are reported to act via other hormone receptors like estrogen receptor, androgen receptors and glucocorticoid receptors. Many studies have reported that they act via the nuclear hormone receptor-mediated pathways. Molecular cross talks with other signaling pathways have also been reported. Steroid hormones have an essential role in lipid storage and disposition of body fat. Estrogen based hormone replacement therapy is prescribed to women at their menopause to remodel their adipose depot. Foetal or neonatal exposure to phytoestrogens may induce

**Figure 2.** *PPAR*γ*-RXR mediated action of obesogens.*

obesity in later stages of life. Well-known phytoestrogen genistein, commonly found in soy-based foods, affects adipose tissue deposition in a dose-dependent and gender-specific manner [46].

Neonatal exposure of DES to female mice led to weight gain in adulthood. However, this effect can be sex-biased. While some EDCs may act directly via cellular steroid receptors by inducing estrogen synthesis, other EDCs may act indirectly. It is established that adipose tissue is a site of estrogen synthesis. The adipocyte cytoplasm contains the enzyme cytochrome P450 aromatase, which plays a vital role in converting estrogen from androgen. It is now reported that several EDCs can impair intracellular aromatase activity [47]. This action may raise intracellular estrogen levels in adipocytes and lead to obesity irrespective of the sexes [48]. It is reported that TBT can directly reduce the activity of the aromatase enzyme in adipose tissue at high doses, leading to reduced estradiol levels and down-regulation of the ER target genes. TBT also has an inhibitory effect on 11β-hydroxysteroid dehydrogenase 2, which leads to reduced inactivation of cortisol. It is believed that increased glucocorticoid levels could influence adipocyte differentiation and regulation of metabolism [40].

Some obesogens, especially the persistent organic pollutants (POPs), act via the ligand-activated transcription factor aryl hydrocarbon receptor (AhR). AhR activates xenobiotic-metabolizing enzyme cytochrome P450s. They can promote adipogenesis indirectly by changing PPARγ expression.

#### **4.3 Other mechanisms**

In some recent studies, researchers found that they are not linked to activation of any nuclear hormone receptors; instead, they followed some novel mechanisms, which make their mechanism of action more complex. Those include epigenetic modifications, impairment of thermogenesis and dysbiosis in gut microbiota. Some of these mechanisms will be discussed in the following sections. Some recent studies correlated COVID-19 pandemic to the obesogenic exposures, that is also being discussed in this chapter.

#### *4.3.1 Epigenetic modifications*

Epigenetics is defined as the study of heritable changes in phenotype resulting from environmentally influenced modifications of genome. Epigenetic modification can alter gene expression during development and cellular differentiation in response to environmental factors such as chemical contaminants. These modifications include DNA methylation at cytosine residues of 5′ to guanine sites (CpG sites), chemically modifying histone proteins and noncoding RNAs interference [49]. DNA methylation was considered a key mechanism responsible for adult diseases with developmental origins [50]. DNA methylation changes are responsible for the transgenerational effects of exogenous exposed individuals to chemicals and nutrition deficits [51]. For instance, the obesogen pesticide TBT induced changes in DNA methylation and histone modification invitro. Various reports have documented the environmental chemicals, including obesogens, led to an epigenetic modification in vivo and obesogen phenotype even in unexposed generations. TBT exposure in 3 T3-L1 mice preadipocytes invitro resulted in increased adipocyte differentiation along with decreased DNA methylation levels. Increased differentiation level towards the adipogenic lineage was observed in adipose-derived stromal cells (ADSCs) isolated from TBT exposed mice perinatally but at the cost of decreased osteogenesis. ADSCs exposed to TBT were associated with increased adipogenesis marker genes, such as PPARγ target gene Fapb4, where methylation level decreased in the promoter region. However, PPARγ mRNA levels were increased, but DNA methylation at its promoter region had no effects [43]. A possible reason for this lack of epigenetic regulation might be that EDC exposure during differentiation process causes DNA histone demethylation. Ultimately, PPARγ, which is under the control of H3K27me3, causes the gene to be promptly up-regulated. Importantly, prenatal exposure to TBT has been recently shown to cause the transgenerational inheritance of adiposity. It remains to be determined whether these transgenerational effects are related to permanent changes in DNA methylation profiles or other epigenetic processes.

#### *4.3.2 Impairment of thermogenesis*

Recent advances found in understanding adipocyte function was the presence of thermogenic brown adipose tissue (BAT) in adult human beings in a dispersed manner, not as found in concentrated discrete depots in human infants. Another discovery of white adipose tissue can also be induced to produce thermogenic fat called beige or brite fat. Increased mitochondria production is responsible for differentiation of both bona fide brown adipocytes and beiging of white adipocytes. This thermogenesis relies on the capacity to dissipate energy in the form of heat through uncoupling of cellular oxidative phosphorylation and ATP synthesis via Uncoupling protein 1 (UCP1) or sometimes through shivering. Some of the evidence has documented how some obesogens impede the production and function of thermogenic adipocytes. For instance, perinatal exposures to DDT in mice have long term-effects on thermogenesis regulation in their female offspring. When female offspring reached up to 6 months of age, they showed reduced energy expenditure & ultimately decreased thermogenesis capacity. However, no change in their physical activity was observed. Thermogenesis impairment was due to the decreased expression of PPAR-γ co-activator 1α (Ppargc1a), a master regulator

for thermogenesis related genes and type 2 iodothyronine deiodinase (DiO2) (the enzyme that catalyzes thyroid hormone T4 to convert into T3 which stimulates BAT thermogenesis) [52]. Secondly, Shoucri and his colleagues [49] found that TBT or rexinoids have inhibited adipocytes' production. Other EDCs increase thermogenesis by changing mRNA and protein levels of UCP-1. Adult mice exposed to PFOA and PFOS through diet (containing 0.02% w/w) for ten days exhibited BAT mitochondria activation for increased oxidative capacity and protein levels of UCP-1, resulting in decreased depots size of adipose tissue. PFOA exposure (80–40 μM) during in-vitro experiments activates UCP1 similarly as fatty acids. These examples indicate how obesogens influence obesity by impairing thermogenesis during the in-vitro and in vivo study. This intriguing area of obesogen epidemic and their mechanism remains to be elucidated. Through their Horizon 2020 programme, the European Union has funded several grants to establish new assays to assess EDCs effects on metabolic-end points and identify those chemicals that affect thermogenesis [53].

#### *4.3.3 Gut microbiota dysbiosis*

The gut microbiome is defined as "the totality of microorganisms, bacteria, viruses, protozoa, and fungi, and their collective genetic material present in the gastrointestinal tract" by molecular biologist Joshua Lederberg. Obesogen exposure could lead to obesity by altering the gut microbiome, a relatively novel mechanism which leads to obesity. It is well understood that obesity is correlated with gut microbiome composition [54]. Some experimental data shows that the transplant of gut microbe from obese mice can induce obesity in lean mice [55]. Conversely, the gut microbiome transplant from lean donors improved the metabolic disorder condition in obese mice [56]. It is evident from several experimental data that many obesogens induce the gut microbiome dysbiosis in zebrafish [57], mice [58] and human [59]. In mice, gut microbial dysbiosis was associated with increased fat accumulation or impaired lipid metabolism after exposure to triphenyl phosphate. Tributyltin exposure induces gut microbiome dysbiosis with increased body weight gain and dyslipidemia in mice [58]. Though, it is not yet apparent whether this metabolic disruption is a result of the gut microbiota dysbiosis or not.

Additionally, some microbial metabolites have also been reported as AhR agonists and antagonists [60, 61], as we are already aware that activating AhR inhibits adipogenesis. In contrast, inhibition of the activity leads to obesity and fatty liver disease. Two basic dietary emulsifiers, carboxymethylcellulose and P-80, were reported to initiate intestinal inflammation and gut microbiota dysbiosis, which led to metabolic disorder and increased body weight in mice [62]. These pieces of evidence suggest that inducing obesity via gut microbiota dysbiosis is possibly a potent mechanism for the obesogens to follow. However, to get more clues, this field needs to be studied further extensively.

#### *4.3.4 Obesogens and COVID-19*

The current outbreak of novel coronavirus has emerged as a worldwide pandemic in the past year, which is related to the Severe Acute Respiratory Syndrome Coronavirus (SARS-CoV) in 2003 and the Middle East Respiratory Syndrome Coronavirus (MERS-CoV) in 2012 [63]. Interestingly, a study in 2003 found a positive correlation between air pollution and extreme SARS in the Chinese population. Patients with SARS from regions with a high air pollution index (API) were twice as likely as those from regions with low APIs to die from SARS [64]. A finding based on US population found that long-term exposure to air

#### *Environmental Obesogens and Human Health DOI: http://dx.doi.org/10.5772/intechopen.96730*

pollution resulted in a 6% rise in cardiopulmonary mortality risk. Some of these pollutants are potent obesogenic [65].

Human studies have even shown nitrogen dioxide (NO2), one of the components of air pollution, is correlated with higher fasting serum lipids among obese individuals, indicating that obesity can worsen the effects of air pollution [66]. Animal studies have also shown that air pollution particles' sensitivity early in life will contribute to increased visceral obesity, insulin tolerance, and inflammation, signaling NO2's function as an endocrine disruptor [67]. Since COVID-19 is similar to SARS in causing respiratory disease, exposure to NO2 can increase the mortality rate of patients with COVID-19. However, future studies are needed to validate this relationship.

#### **5. Transgenerational effects**

One of the most intriguing results in EDCs field came when a series of reports were published by Skinner and colleagues showing EDCs, including DDT and MTX, induce transgenerational obesogenic effects. During F1 generation, prenatally exposed individuals with anti-androgenic fungicide vinclozolin or estrogenic pesticide MTX were associated with disease in various organs in their F4 generation [68]. Similarly, when pregnant mice (FO generation) were exposed to environmentally relevant doses (nM) of TBT through drinking water, then effects on obesity were observed in F1-F3 descendants of exposed animals [69]. Notably, in a similar experiment, the pharmacological obesogen, Rosiglitazone, which can activate PPARγ, could not produce the same transgenerational obesity effects suggesting that different pathways in addition to PPARγ were required to generate transgenerational phenotype [69].

In addition to TBT effects on obesity, Skinner lab has shown several environmental chemicals such as plasticizer (BPA, DEHP, DBP) [70], pesticides MTX [71], a mixed hydrocarbon mixture (jet fuel JP-8) [72] and the widely used pesticide DDT [73], induced transgenerational obesity in a rat model as observed in F3/ F4 offspring of ancestral prenatal or perinatal obesogen exposed-FO individuals [71–73]. Although molecular mechanisms underlying transgenerational inheritance of obesity are currently controversial, researchers belonging to the EDC field believe that these obesogen effects are inherited in an epigenetic manner. This point has got stronger resistance in the genetics sphere [74].

#### **6. Epidemiological evidence of obesogens**

#### **6.1 Human cohort studies**

Epidemiological studies are of considerable significance for the association of disease effects with exposure to obesogens. Few cohort-based studies are available to date on the effect of obesogens in human populations. Since a considerable amount of evidence indicate that prenatal exposures predispose patients to obesity, epidemiological research concentrates on obesogenic measurements throughout pregnancy. Increase in child adiposity in multiple birth cohorts was associated with prenatal exposure to PFAS. At the same time, sexual dimorphism was sometimes linked with it [75–79]. A metapopulation analysis, including ten cohorts, suggests a 25% and 0.1 unit increase in weight and BMI, respectively, per ng/ml of PFOA concentration in maternal blood [80].

A research found that rising concentrations of maternal urinary phthalate during gestation doubled the risk of the offspring becoming overweight or obese [81]. Cohort research on the impact of prenatal BPA exposure has also been correlated with increased waist circumference, BMI, and risk of obesity [82]. Studies of prenatal exposure to phthalates and bisphenols have not shown a consistent association with measures of childhood adiposity compared to studies of prenatal exposure to PFAS [83]. Two studies on the American population showed an association between serum concentrations of PFAS and weight gain irrespective of sexes [84]. PFAS, particularly perfluorooctane sulfonate (PFOS) and perfluorononanoic acid (PFNA), were linked with alteration in metabolic rate [85].

Few studies have explored the longitudinal impacts on postnatal growth of prenatal exposure to other chemicals. Evidence risen over the past five years indicates that exposure to phthalates leads to adult weight gain, with most research conducted in women. Some studies by the Women's Health initiative reported a strong correlation between urine concentrations of phthalate metabolites and weight gain [86]. Again, it is to be considered that the effect of a single chemical mostly reflects the epidemiological studies conducted. However, naturally, obesogens ploy cumulative effect as mixtures. The WAT is the depot of obesogens in the human body. More studies should be designed to estimate the accumulative effect of mixtures in future.

#### **6.2 In vitro models**

In vitro models have several advantages over other model systems. Taking human cells lines for the study can be of great significance considering the physiological relevance. For screening new chemicals for potential obesogenic properties, in vitro studies are generally conducted before animal models. Several cell lines are used to study the obesogenic impacts of several compounds. Among the in vitro models, mouse embryo pre adipocyte 3 T3-L1 has been used extensively to check the effects of obesogens like TBT [87], BPA [88], BPS [89], genistein [90], phthalate [91], nonylphenol [92] and so on. Other cell lines include C2C12 (mice muscle cells) [93], HELA (human cervical cancer cells) [93], HEK293C (human embryonic kidney cells) [94], HepG2 (human liver carcinoma cells) [95], hASCs (human adipose-derived stem cells) [96], C57BL/6 (mice bone marrow stromal cells) [97], hESCs (human embryonic-derived stem cells) [98] etc.

#### **6.3 In vivo models**

Though animal models are not recommended to study certain chemicals' obesogenic potential, they do not mimic the human physiological systems. Still, in vivo model systems have certain advantages over in vitro systems as whole-body kinetics and systemic effects can be studied using animal models. Complex linked pathways involving multiple organs, including adipose tissue, liver, pancreas, muscle, brain, etc., regulate metabolism and weight. In understanding the role of chronic inflammation and hormone interference, in vivo experiment is particularly relevant. The most widely used animal model for the study of obesogens is rodents. Multiple obesogens including TBT [69], BPA [99], triphenyltin [100], DEHP [101], DES [102], polycyclic aromatic hydrocarbons, DDT, and nicotine, have been defined as murine models. Mice are identical biologically and anatomically to humans and share many common diseases. It is incredibly useful for diseases with an inflammatory condition, such as obesity, as animal models can mimic complex inflammatory responses. A transgenic model like obese or lean bodied mice can also be created by manipulating required genes. Other commonly used in vivo models include rats [103], zebrafish [104] and drosophila [105]. Many insights into possible obesogens and various modes of action were provided using in vivo models to investigate

endocrine disruption. They may not replicate human physiology, as discussed earlier. Mice exposed to a specified amount of one particular molecule over weeks sometimes does not reflect a chronic variable exposure in humans to multiple chemicals over the years. In detecting obesogens and discerning mechanisms of action, animal models play an essential role. However, they should be combined to draw the most reliable conclusions with knowledge from in vitro studies and epidemiological studies.

#### **7. Strategies for change and future directions**

The obesity epidemic first continues in the US and afterwards expands worldwide; therefore, it becomes a dire need to understand the predisposition and related disorders' mechanisms. It becomes of utmost importance to study the extent to which the obesogen exposure influences obesity in humans and establishes the risk factors related to obesity. The risk factors include oxidative stress, inflammation, disrupted circadian rhythms, mitochondrial dysfunction and dietary composition. These interactions may be critical in the effects of obesogen exposure. Evidence documented in the obesogen research area shows that their effect mainly depends on the level and timing of exposure, especially critical windows of exposure during fetal development. Hence, it is crucial to reduce or avoid exposure to obesogens, specifically during pregnancy. However, there is no technique to determine if the individuals have been exposed to any obesogens during their development. It will be a "Holy grail" to identify biomarkers of exposure in obesogen research and establish links among obesogen exposure and other factors related to obesity. The obesogen hypothesis opened a new field into obesity by linking EDCs research with developmental disease origin. The obesogen hypothesis is still in the dearth of research. It requires more studies in the mechanism, developmental time windows and diet interaction. The effects of obesogens are related to epigenetics.

However, we still need more research to understand the mechanism and how the effects get transmitted to forthcoming generations. For instance, how does the obesogen exposure of pregnant Fo female mice lead to obesity in upcoming F3 and F4 unexposed males? There is an extreme lack of data on how obesogen exposure programs adipose tissue dysfunctional that could readily store but not mobilize fat. The obesogen sphere is almost 15 years old only. Much has been studied related to potential effects of EDCs and obesogens. The most substantial evidence for chemical obesogens existence may be the variety of pharmaceuticals that have the side effects of making patients obese. Several international and national workshops have been held to understand the potential role of EDCs in obesity and related metabolic disorders [53]. Thus, various policies and strategies should investigate the magnitude of environmental obesogenic pollutants on the obesity epidemic and the regulatory actions required on such chemicals to improve public health.

#### **8. Conclusion**

The majority of evidence that indicates the role of EDCs in driving obesity provides a mechanistic explanation of the obesity epidemic and a management strategy. The role of exogenous chemicals in growing rates of obesity through gene expression regulation (such as PPARs), hormone changes, and inflammation is supported by ample evidence. While overeating, combined with lack of exercise,

is undoubtedly a significant contributor to the increase in obesity that can be addressed by decreased calorie intake and increased exercise, it may be that reducing exposure to obesogenic EDCs may also contribute to reducing obesity in the population, especially during the early stages of life. More knowledge of obesogenic pathways will improve prophylactic and therapeutic strategies. The extensive exposure of the human population to so many EDCs with obesogenic action needs evaluation. In vitro models are useful screening devices for detecting and testing obesogenic mechanisms, notably, changes in gene expression or molecular pathways. Improvements to these models will improve human extrapolation in vitro to in vivo as well. However, animal models remain a valuable and typically physiologically precise method for studying obesogenic inter-organ pathways, including hormone interference and inflammation. More epidemiological studies should be made to confirm in vitro and in vivo animal models and provide unparalleled insight into human obesogen exposures and effects. Integrating the data collected from all three of these model systems would result in better-informed choices of compounds that can be used to replace obesogens in food production, packaging, etc. It will, essentially, reduce the economic burden of obesity.

#### **Acknowledgements**

All authors thank Banaras Hindu University, India, for providing necessary resources and support to write the present chapter with the support from University Grant Commission (UGC)-Junior Research Fellowship to AM, PG and AS. This work received no external funding from any agency.

#### **Conflict of interest**

The author declares that there is no conflict of interest.

#### **Author contributions**

All authors listed have made a substantial contribution to this chapter. Moreover, special thanks to PG for writing some sections and proofreading the whole manuscript. Thanks to RKS for reviewing the manuscript before the final submission.

*Environmental Obesogens and Human Health DOI: http://dx.doi.org/10.5772/intechopen.96730*

#### **Author details**

Archisman Mahapatra, Priya Gupta, Anjali Suman and Rahul Kumar Singh\* Department of Zoology, Molecular Endocrinology and Toxicology Lab (METLab), Institute of Science, Banaras Hindu University, Varanasi, India

\*Address all correspondence to: rks.rna@gmail.com

© 2021 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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#### **Chapter 4**

## Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal, Cardiovascular and Hypertensive Diseases

*Rafael Moreno-Gómez-Toledano, María I. Arenas, Sandra Sánchez-Esteban, Alberto Cook, Marta Saura and Ricardo J. Bosch*

#### **Abstract**

Bisphenol A (BPA), an endocrine disruptor involved in synthesizing numerous types of plastics, is detected in almost the entire population's urine. The present work aims to estimate daily exposure to BPA by systematically reviewing all articles with original data related to urinary BPA concentration. This approach is based on human pharmacokinetic models, which have shown that 100% of BPA (free and metabolized form) is eliminated only in a few hours through urine. Several extensive population studies and experimental data have recently proven a significant association between urinary excretion of BPA and albuminuria, associated with renal damage. Our team's previous work has shown that low-dose BPA can promote a cytotoxic effect on renal mouse podocytes. Moreover, BPA administration in mice promotes kidney damage and hypertension. Furthermore, preliminary studies in human renal cells in culture (podocytes) strongly suggest that BPA might also promote kidney damage. Overall, the present review analyzed BPA exposure data from mammalian cell studies, experimental animal models, and several human populations. Studying principal cohorts calculated the exposures to BPA globally, showing a high BPA exposure suggesting the need to decrease BPA exposure more effectively, emphasizing groups with higher sensitivity as kidney disease patients.

**Keywords:** bisphenol A, systematic review, human, urine, estimated daily intake

#### **1. Introduction: brief historical overview**

Bisphenol A is the perfect example of the double edge of industrial development. On the one hand, thanks to BPA, we have countless plastic objects with excellent physical properties at low prices; on the other hand, increasing exposure to this kind of xenobiotic compounds could be a severe health risk to the general population.

BPA is a phenolic compound widely distributed due to its multiple uses as an additive and plasticizer in plastic polymers' manufacture [1]. This compound can be found in various everyday items, such as food containers, toys, dental supplies, electronic devices, and even clothing [2–6].

The BPA problem presents a particular and curious situation: BPA is a compound whose properties as an estrogen modulator were already determined 84 years ago by medical researchers at the University of London [7], but its use increased substantially last decades. The discoverer's idea was to commercialize a compound that could treat female pathologies. Finally, they succeeded with Diethylstilbestrol, a substance with much greater potency than BPA, and was introduced in the 1940s [8].

It took about 50 years since the Russian chemist Dianin synthesized it in 1891 [9, 10] until the BPA began to be used in the industrial manufacturing of epoxy resins. Still, due to its incredible versatility, BPA quickly achieved great importance in the American industry. In the mid-1970s, the BPA was considered a part, directly or indirectly, of all major US industries [8]. In parallel, Schnell's contributions in 1956 demonstrated BPA's potential role in producing polycarbonates [11, 12]. Due to its unique combination of physical properties, this type of compound has had a significant impact on the world industry, as have epoxy resins. Today they are still used in numerous applications, such as in the automotive or LED sector [12]. In fact, there is a tendency to increase its consumption in the coming years, as can be seen in the Asian market, where there has been a substantial increase in the demand for polycarbonates in the last ten years [13]. It is expected to continue growing in the years to come, as observed in the American market [14].

#### **2. Novel role of BPA in renal, cardiovascular and hypertensive diseases; latest discoveries**

#### **2.1 BPA in the renal system**

BPA is a compound widely studied for its estrogenic properties within the field of fertility and sexual organs. However, other organs, such as the kidney and liver, may have the highest exposure ranges. In the kidney's case, BPA concentration has been positively correlated with a greater predisposition to kidney pathologies [15–17] or clinical signs associated with kidney diseases, such as increased albuminuria or decreased glomerular filtration rate [18–21].

Our group has worked on BPA's possible action on the renal system in recent years, using different cell and animal models. The first steps were carried out on a renal cell line of immortalized mouse podocytes. It was possible to observe how the chronic treatment of BPA exerts a cytotoxic effect on the cells. The administration of 10 and 100 nM doses for nine days exerted loss of cell viability and increased apoptosis (as assessed by MTT and TUNEL, respectively). These effects were accompanied by an increase in the synthesis of molecules classically involved in the pathogenesis of glomerulosclerosis, such as the cyclin-dependent kinase inhibitor p27kip1, the TGF-β system, and collagen IV. Furthermore, in these cells, BPA reduced the synthesis of nephrin and podocin, proteins of the filtration slits involved in proteinuria and podocyte survival mechanisms. As would be expected from these in vitro results, the kidneys of animals treated with BPA developed hypertrophy, hyperfiltration, and proteinuria. Along with the increased renal expression of p27kip1, TGF-β, and collagen IV, mesangial expansion and a decrease in the number of podocytes due to apoptosis were also seen. Electron microscopy showed hypertrophy of podocytes and pedicles. It should be noted that even when animals treated with BPA did not develop hyperglycaemia, their kidneys showed

*Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal… DOI: http://dx.doi.org/10.5772/intechopen.96309*

structural and functional changes similar to those that occur in the initial stages of diabetic nephropathy (DN) (**Figure 1**) [22, 23].

Secondly, the possible effects of BPA on an immortalized human podocyte cell line were explored. We observed that BPA promotes a novel type of podocytopathy characterized by an impairment of cell adhesion by altering adhesion and cytoskeleton proteins' expression.

By using transcriptomics, proteomics, western-blot, and immunocytochemistry, it was possible to determine that BPA at low doses promotes a reduction in the expression of numerous structural or adhesion proteins, such as tubulin, vimentin, podocin, cofilin-1, vinculin, E-cadherin, nephrin, and VCAM-1, as well as an increase in the expression of proteins that negatively participate in adhesion mechanisms, such as Tenascin-C [24].

Since podocytes do not replicate in adults, the resulting podocytopenia after the urinary loss of podocytes might promote glomerulosclerosis. Collectively all available data suggest that BPA could participate in the pathogenesis and progression of renal diseases. It is essential to mention that these experimental results are supported by epidemiological studies conducted in the populations of New York [25], Shanghai [19],

#### **Figure 1.**

*Notable evidence from cell and animal models. A) TUNEL technique performed in mouse podocytes. Note the significant increase in the number of apoptotic cells in the cells treated with BPA. B) Hypertensive effect of BPA administered to animals in drinking water. C) Reduction in the number of glomerular podocytes (labelled with WT-1) of the mice treated with intraperitoneal (IP) BPA. D) Increased number of apoptotic kidney cells (TUNEL) of mice treated with IP BPA. E) Adhesion assay in human podocytes. A reduction of up to 50% was observed in cells treated with BPA. F) Podocin immunocytochemistry in human podocytes. A significant loss of labelling is evidenced in cells treated with BPA. \* or @ means p-value < 0,05; \*\* or @@ means p < 0,01; \*\*\* or @@@ means p < 0,001; \*\*\*\* or @@@@ means p < 0,0001. Figure made with our own results published in the Journal of Cellular Physiology [22], FASEB Journal [27] and Scientific Reports [24].*

and Seoul [26], which describe an association between human exposure to BPA and an increase in proteinuria and hypertension.

#### **2.2 BPA in the cardiovascular system**

Further studies demonstrated that animals treated with BPA developed arterial hypertension and endothelial dysfunction in a dose-dependent manner (**Figure 1**). Microarray analysis of gene expression in murine endothelial cells treated with BPA demonstrated the activation of genes involved in vascular regulation, such as angiotensin II and calcium-calmodulin kinase II (CaMKII). This event was subsequently observed in vivo as well. The activation is responsible for the endothelial dysfunction and hypertension induced by BPA, given that CaMKII activation promotes the enzymatic uncoupling of endothelial nitric oxide synthase. This phenomenon leads to oxygen free radicals' production instead of nitric oxide, a primary vasodilator, and endothelial protector. Moreover, this increased production of oxygen free radicals indicates that BPA, and inducing hypertension, could participate in vascular damage mechanisms and atherosclerotic lesions' progression [23, 27]. Besides, recent data demonstrated the cardiotoxic effect of BPA by a mechanism that involved activation of the RIP 3-CamKII necroptotic pathway leading to endothelial cell death. Decreased endothelial barrier function and weakening of the coronary vascular wall in the setting of hypertension may cause ventricular hemorrhages, cardiac and lung congestion, which ultimately led to heart failure [28].

#### **3. BPA exposure in the general population. Identification of groups with higher exposure**

#### **3.1 Pharmacokinetics of BPA**

The heterogeneous distribution of BPA results in the ability to enter the body in multiple ways. The main entry route is considered oral, through the ingestion of food or beverages containing BPA [29, 30]. However, there are other routes like inhalation (air or dust) [31–34], dermal (cosmetics, thermal tickets) [35–37], and it has even been hypothesized with the ocular [38] and sublingual routes [35, 39]. It is estimated that between 85 and 100% of the BPA ingested can be absorbed through the intestine. Thanks to its capacity to cross biological barriers, it has been observed that BPA has the potential to distribute itself through any fluid and biological tissue, even crossing the transplacental or blood–brain barrier [29, 40, 41]. In the case of the dermal route, it has been determined that the ability of BPA to enter the body is lower, with percentages less than 10% [42, 43]. For its part, the sublingual route (of great importance in the elements used in dentistry) seems that it could become more efficient than the intestinal entry [39].

BPA's metabolism is marked by phase II reactions, biochemical mechanisms capable of modifying its structure to facilitate its excretion [44]. BPA is metabolized towards glucuronidation or sulfonation in the intestine and the liver [41, 45, 46], but the metabolic capacity can be seriously reduced in diseases such as obesity or diabetes [47]. Glucuronidation is the majority reaction, mediated by uridine diphosphate glucuronosyltransferase (UGT) [44, 48]. It has also been suggested that a part of the BPA that reaches the intestine could be degraded to p-cresol by the intestinal microbiota, thus generating uremic toxins [49]. Another possible route studied has been hydroxylation to catechol, followed by a transformation to o-quinone. This route, like the previous one, can generate toxicity associated with oxidative stress [50].

*Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal… DOI: http://dx.doi.org/10.5772/intechopen.96309*

Pharmacokinetic studies in rodents have determined that BPA is excreted in urine and feces [29, 51, 52]. It has been observed that BPA is excreted exclusively through the urine in non-human primates and humans [41, 53, 54]. This phenomenon makes it much easier to make a rough estimate of the degree of exposure by BPA's urinary quantification. The inter-species differences observed are attributed to a possible higher enterohepatic recirculation in rodents [29, 51, 52]. However, there is evidence that contradicts this hypothesis [55].

#### **3.2 Calculation of BPA exposure in the general population**

As mentioned above, thanks to pharmacokinetic studies in humans [41, 53, 54], it is accepted that 100% of BPA is eliminated via the urinary tract, which can be used to determine the degree of daily exposure to this compound quickly. For this reason, we proceed to evaluate the question of the degree of global exposure through a systematic review of principal cohorts in the world. To estimate human exposure to BPA, we first collected data published by one of the world's largest cohorts: the National Health and Nutrition Examination Survey (NHANES). NHANES is a survey research program conducted by the US National Center for Health Statistics (NCHS), with more than 72,000 patients studied between 2003 and 2016 [56].

After extracting all the data and unifying them, 18,244 urinary BPA concentrations were obtained. A non-parametric distribution was obtained after performing


*Note that two results have been included with a reduced sample size to include Oceania and Africa. Meaning of abbreviations: K, kids; A, adults; GM, geometric mean (or corrected median). \*Own study corresponds with NHANES 2003 – 2016 cohort.*

#### **Table 1.**

*Representative data from the main cohorts in the world.*

the normality tests, for which the geometric mean (GM) was calculated, obtaining a result of 1.77 ng/ml. A systematic review of urinary BPA was then carried out to select from among all the publications with the most representative cohorts from each continent and the largest number of people. Using the keywords: Bisphenol AND (urine OR urinary) in the reference search engines Pubmed and Web of Science, a total of 999 and 2,025 results were obtained, respectively. Once the duplicates were eliminated, a total of 2,414 publications remained. After screening by title/abstract, a total of 756 publications were selected. Finally, after reading in-depth, 447 articles were selected whose pages describe urinary concentrations of BPA in some population groups, either general or specific, such as patients with various pathologies, pregnant women, the elderly, or workers subjected to occupational exposure. All data from the 447 academic articles were collected and analyzed carefully. According to the country, population group, and sample size, the primary world cohorts were selected from all of them, obtaining 16 cohorts whose sample sizes exceed 1000 individuals from America, Asia, and Europe (**Table 1**). A result of Oceania and another from Australia was also included due to representability. All of them expressed the concentration of BPA in ng/ml except one of them, which expressed it in μg per gram of creatinine (μg/g creat.) [71]. Therefore, it was modified by calculating the average creatinine concentration in adults using the NHANES cohort's data and the other two major cohorts, KoNEHS and CHMS [74, 75].

We consider that in the study of urinary BPA, where the results follow nonparametric distributions, the values that should be analyzed would correspond to the GM or the median. To determine if both values can be unified, they were examined using linear regression, observing that they were always in the same range, and the variation between them was relatively small. The equation of a line was **Y = 0.9855 \* X, and R<sup>2</sup> = 0.9919**. For this reason, the decision was reached to


#### **Table 2.**

*Higher urinary BPA values determined according to the geometric mean (GM), median, 95th percentile (P95), or maximum value (MAX).*

*Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal… DOI: http://dx.doi.org/10.5772/intechopen.96309*

use the GM, preferably, but if it was not recorded, use the median corrected with the equation obtained.

Next, the urinary BPA was averaged, considering each cohort's sample size, obtaining a final result of **1.55 ng/ml** (with a sample size of 57,537 individuals). Once the average concentration determined in the general population's urine has been established, the next step will be carried out on the highest values found in the systematic review to determine interest groups to study BPA exposure.

As reflected in **Table 2**, it is clear that workers subjected to occupational exposure are the ones who are likely to find a more significant entry of BPA into their bodies. The highest values observed, both of the median and the maximum value (MAX), correspond entirely to people subjected to occupational exposure, such as workers in the plastics industry. The highest GM value stands out, as it corresponds to the general Iranian population. An in-depth study would be necessary to be able to discern the problem that underlies this study area. Interestingly, unusually high 95th percentile values can also be seen in pregnant women and intensive care patients. It is likely that consumable medical supplies, such as catheters or hemodialyzers, could increase BPA exposure due to their plastic composition. Bearing this in mind, and in keeping with the discoveries described in basic research, the study and analysis of BPA exposure in patients undergoing hemodialysis is crucial.

#### **4. Systematic review of BPA exposure in hemodialysis patients**

After describing the latest advances in the BPA-kidney paradigm investigation, the need to include kidney patients as a group of special vulnerability to exposure to BPA is evident. Thereof, there is a point of convergence in the final stages within the different pathologies or stages: the need for dialysis due to the kidney's reduced functionality. Interestingly, there is evidence that the use of surgical medical equipment can increase exposure to compounds such as BPA due to the composition of its materials. Therefore, we will analyze the urinary concentration of BPA in patients undergoing hemodialysis procedures to estimate the daily exposure to which they are subjected. The systematic review methodology was used again, using the keywords: bisphenol AND (dialysis OR hemodialyzer OR hemodiafiltration OR hemodialysis OR dialyzer).

Thirty-eight results were obtained in Pubmed and 50 in Web of Science. After eliminating duplicates, a total of 66 documents were obtained. Once the first screening by title/abstract was done to look for BPA concentrations in patients undergoing hemodialysis, a total of 20 publications were accepted. After carefully studying the text, ten publications were selected1 . Of these, only 1 quantifies the urinary BPA concentration in dialysis patients [87] and 9 in serum [15, 86–94].

The publication by Schöringhumer et al., which quantifies urinary BPA, obtains concentrations between 0.4 and 2.6 ng/ml within the same range as the general population [87], equivalent to 1.75–11.39 nM. In general terms, low exposure would be considered. Still, considering the in vitro model results and the patient's pathology, it could pose an added risk for kidney disease evolution. In the case of publications that study BPA in serum, some show values similar to those observed in the general population's urine. Among them, we can find the publications of Kanno et al. (5.3 ± 0.3 ng/ml), Murakami et al. (values between 1.48 ± 1.41 and 6.62 ± 3.09), Sajiki et al. (values between 0.179 ± 0.263 and 0.642 ± 1.443), Shen et al. (1.01) or Turgut et al. (5.57 ± 1.2) [88, 90, 92–94]. Higher values have the publications of Quiroga et al. (high flux hemodialysis: 7.5 ± 3.5; online hemodiafiltration: 6.7 ± 2.5)

<sup>1</sup> Those publications without relevant data, reviews, and conference communications were discarded.

and Krieter et al. (10 ± 6.6) [15, 91]. Finally, Bosch-Panadero et al. and Mas et al. describe serum BPA values in patients undergoing conventional dialysis that range from 52.73 ± 60.6 to 163.03 ± 155.84. Also, they quantify serum BPA concentrations in patients undergoing online hemodiafiltration from 8.79 ± 7.97 to 23.42 ± 20.38) [86, 89]. These high values would be equivalent to 230.98–714.14 nM in the case of conventional dialysis and 38.50–102.59 nM in online hemodiafiltration.

#### **5. Systematic review of occupational exposure to BPA**

The alarming data described in the previous pages denote the need to study occupational exposure. To this end, we proceeded to use two academic reference search engines, Pubmed and Web of Science, using the following keywords: Bisphenol AND (workers OR occupational exposure OR exposure workplace), obtaining a total of 658 publications (once repeated results were eliminated). Of all of them, 25 publications were adapted to the search. Only publications with urinary BPA (or blood) concentrations were selected in workers with high exposure or themselves before and after their work shift. Of the 25 studies selected for their affinity with the topic of interest, we can distinguish three subgroups: In the first (G1)2 , BPA concentrations can be observed well above the average, and with significant differences between the study groups. In the second group (G2)<sup>3</sup> , there are concentrations higher than the mean in a range closer to it, while in the third group (G3), the range of concentrations is within the range of values of the general population.

#### **5.1 G1: extremely high BPA concentrations**

From a quantitative perspective, within G1, the most interesting publication is Liu et al. [95]. It compares BPA concentrations in people with potential occupational exposure versus controls, obtaining substantially different values. The median values (interquartile range, IR) between exposed workers vs. controls are 685.9 (43.7–3671.8) vs. 4.2 (0–15.9) μg/g creat. Other equally interesting values are Tian et al. [78] and Song et al. [77]. Firstly, they determine geometric mean (GM) values (standard deviation, SD) between exposed subjects vs. control of 158.41 (17.92) vs. 0.84 (6.53) μg/g creat., reaching in the second publication, the values of 199.13 (19.65) vs. 0.77 (6.33) μg/g creat. Song et al.'s publication determine the highest maximum urinary BPA concentration, reaching the value of 264,219.38 μg/g creat., (264.22 mg/g creat.).

The next publications to consider are two by Hines et al. [79, 96], where they study exposure to BPA in different factories before and after the work shift. In them, essential differences can be appreciated, showing, to cite an example of each article, a GM (SD) in pre-shift vs. post-shift of 6.2 (4.3) vs. 130 (10) μg/sample or 26.6 (5.74) vs. 178 (6.2) μg/g creat. These groups show arithmetic mean (SD) values of 15 (22) vs. 2300 (5800) μg/sample and 115 (252) vs. 812 (2330) μg/g creat. The maximum BPA value is also very striking, reaching 32,900 ng/ml. We will continue with the study of publications with high BPA values, Xiao et al. [97], and the two publications by Li et al. [98, 99]. They show differences between exposed workers vs. control, showing medians of 101.94 vs. 0 ng/ml of serum in the first case and

<sup>2</sup> All those publications with geometric means or medians greater than 50 (ng/ml, μg/g creat., μg/urine sample, or ng/ml of plasma) have been selected.

<sup>3</sup> We have selected those publications with values of geometric means/medians lower than 50 and higher than 8 (at least four times above the global mean).

*Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal… DOI: http://dx.doi.org/10.5772/intechopen.96309*

57.9 vs. 1.2 and 53.7 vs. 1.2 μg/g in the other two. Finally, it is important to mention the works of He et al. [80] and Wang et al. [84]. In the first, they find pre-shift vs. post-shift differences of 84.6 vs. 111 μg/g creat. (median) and 4630 vs. 5400 μg/g creat. (AM). In the second publication, they quantify urinary BPA concentrations in workers of an epoxy resin factory, with a GM (SD) of 55.73 (5.48) and a maximum value of 1934.85 ng/ml.

#### **5.2 G2: elevated BPA concentrations**

Within G2, where the concentrations are not so high, articles such as those by Li et al. [100] or Miao et al. [101, 102]. In them, differences between exposed vs. control are determined, observing medians (IR) of 38.7 (6.3–354.3) vs. 1.4 (0.0–17.9) μg/g creat. in the first case, AM (SD) of 36.23 (7.69) vs. 1.38 (6.89) μg/g creat. in the second, and GM (95% confidence interval, CI) of 22.2 (12.4–39.8) vs. 0.9 (0.7–1.1) μg/g creat. in the latter. The same pattern can be observed in the work of Ndaw et al. [85], where higher values are observed in cashiers exposed to thermal tickets vs. controls, determining a GM (SD) of 8.58 (2.83) vs. 3.52 (2.35). For their part, Zhuang et al. [103] carried out a slightly different approach since they determined differences between workers of an epoxy resin company with a working time greater than five years versus those in the company for less than five years. The median values observed reflected a significant increase in workers with a longer working time (27.18 vs. 9.73 ng/ml serum). Finally, the work of Heinälä et al. [104] is also included in this group, where the pre-shift vs. post-shift urinary concentration is studied, quantitatively highlighting the GM of the heat-sensitive paper producing company, 18.7 vs. 39.4 ng/ml, or from the liquid paint producer, 4.6 vs. 10.3 ng/ml of urine.

#### **5.3 G3: "normal" range but with significant differences**

The third group, G3, despite being in the range that we have determined as general, corresponding to the majority of the population, also presents interesting differences. Among them, the works of Zhou et al. [105] and Kouidhi et al. [106] stand out. Their comparison between exposed subjects vs. controls found values corresponding to the median of 3.198 vs. 0.276 ng/ml serum in the first and 3.81 vs. 0.73 ng/ml urine in the second. The same study line is the oldest academic article of the review, published by Hanaoka et al. in 2002 [107]. They determined very few differences between workers in the bisphenol diglycidyl ether (BADGE) industry vs. controls, with medians of 1.06 vs. 0.52 μmol/mol creat. Similarly, He et al. [108] determine few differences between exposed workers and their families, determining a GM of 1.41 ng/ml in exposed men, compared to 0.58 in their women or 0.78 in their children under 20 years of age. Waldman et al. [109] and González et al. [110] also show low GM values. The first measures BPA's urinary concentrations in firefighters, engineers, captains, or battalion commanders, determining a GM of 1.58 ng/ml. In the second, they determine BPA's concentration in workers of an incinerator of hazardous waste, determining a GM of 0.68 in men and 1.2 ng/ml in women. Thayer et al. [111] and Lee et al. [36] carried out two publications focusing on cashiers exposed to thermal tickets. The first determines GM (SD) in pre-shift vs. post-shift cashiers of 1.89 (3.63) vs. 2.76 (3.53) μg/g creat., being 1.25 (1.79) in controls that do not work as cashiers. The second publication finds subtle differences only in those cashiers who do not wear gloves, observing GM values pre- vs. post-shift of 0.4 vs. 0.9 ng/ml in cashiers without gloves, and 0.44 vs. 0.49 ng/ml in tellers with gloves. Finally, it remains to mention the work of Hehn et al. [112], in which analyzing the data from the American health program NHANES according to the possible potential exposure. They determine GM values in women with probable vs. unlikely exposure of 5.45 vs. 2.16 ng/ml, thus as of 2.85 vs. 2.59 ng/ml in men's case.

#### **6. Tolerable daily intake (TDI); calculations and extrapolations**

Tolerable Daily Intake (TDI) is "the maximum amount of a contaminant which can be eaten every day over a whole lifetime without incurring appreciable risk to health" [113]. Currently, the European Food Safety Authority (EFSA) estimates it at 4 μg/kg BW/day [31]. The TDI calculated by EFSA is based on the studies of Tyl et al. [114], in which the concentration limit at which no adverse effects were observed, NOEL or NOAEL, was determined. They used concentrations from 0.03 to 50 and 600 mg/kg BW/day (0.018–3500 ppm) in mice of different generations. They only observed renal effects (increase in organ weight) at the highest dose (600 mg/kg BW/day), thereby determining the NOEL at the next lower dose they used, which corresponds to 50 mg/kg BW/day. Thus, based on the renal NOEL/ NOAEL and due to the presumption of limitations in the use of the parameter, the EFSA calculates the equivalent "Benchmark dose" (BMD). The equivalent concentration, in which it is estimated that there is an alteration in kidney weight in 10% of the treated animals, is 9 mg/kg. After applying a correction factor to estimate the equivalent dose in humans, a concentration of about 600 μg/kg is obtained. Finally, an uncertainty factor of 150 is applied to obtain the final result of 4 μg/kg BW/day [31].

To determine if the population is exposed to a high or low BPA concentration, the estimated daily intake (EDI) must finally be calculated. Exposure levels are expressed as a mass (nanograms or micrograms) per kg of weight per day. For this reason, it is necessary to multiply the urinary concentration of BPA in ng/ml by the average volume of urine (in ml) excreted per day and divide this number by the average weight measured in kilograms (reference values extracted from academic literature [115]). When taking the average value between adult men and women, a value of 1400 ml per day is obtained. The publication itself also shows the reference values for body weight, expressed in kg. When taking the same average as that applied to the urinary volume, adults' average weight would correspond to a value of 66.5 kg. In this way, as reflected in **Table 3**, the main EDIs were calculated.


*Note that the maximum occupational exposure value reaches 5.66 mg/kg BW/day (1000 times higher than TDI). Abbreviations: GM, geometric mean; AM, arithmetic mean; MAX, maximum value; P95, 95th percentile.*

#### **Table 3.**

*Most relevant values in the systematic review.*

*Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal… DOI: http://dx.doi.org/10.5772/intechopen.96309*

#### **7. Discussion**

In the first place, an interesting element to consider resides in the pharmacokinetics models since they add modifications to BPA in order to determine it efficiently and without contamination by HPLC. After the first model made by Volkel et al. [41], where they used d(16)-bisphenol A, successive authors have emulated this methodology in order to accurately measure the pharmacokinetics of administered BPA [54, 55, 116–119]. However, deuterium modification of drugs is used today to reduce toxicity by redirecting metabolic pathways [120]. Perhaps the possibility that not all BPA is excreted in urine should be reconsidered with this in mind. We know that mice excrete BPA in feces; however, there are no publications in the literature that quantify BPA in human feces, although the presence of microplastics in them has recently been demonstrated [121].

Secondly, since it is described that BPA is a hydrophobic molecule, but with slight aqueous solubility and with the capacity to cross all types of biological tissues [29, 41, 122], it is possible their bioaccumulation in the organism. To do this, Richard W. Stahlhut's team determined BPA concentrations as a function of fasting time. Surprisingly, BPA levels did not decrease rapidly with fasting time, suggesting that there may be non-food exposure or bioaccumulation in body tissues [123].

Thirdly, another critical element is the possible non-monotonic effects of BPA on various organs and tissues [124–126]. This non-monotonicity can significantly affect at low concentrations, below the current TDI, in the same way that it has been shown to happen with certain hormonal stimuli. In her review, Vanderberg [124] determined that non-monotonic dose–response curves (NMDRCs) are typical in the literature related to BPA, occurring in greater than 20% of all experiments and at least one endpoint in more than 30% of all studies examined [124]. Going a little deeper into the non-monotonic effects works such as that of Angle et al. demonstrate the existence of multimodal dose–response curves [127]. Recent data suggest that the non-monotonic effect of BPA could depend upon the target tissue. In our studies in mice, we observed that while BPA induces hypertension in a dosedependent manner, it affects renal podocytes in a classical non-monotonic response curve [22, 23, 27]. In multimodal curves, increases and decreases are observed, and variations in the maximum response depending on the type of tissue [127] may further complicate the correct assessment of BPA's presumed safety concentrations currently found in the population.

Throughout this chapter, an average urinary BPA value for the general population has been determined using a systematic methodology to serve as a reference. Similarly, the analysis of the different statistical parameters shown in the publications determined population groups of special interest, such as workers with occupational exposure, pregnant women, or intensive care patients. With the latest discoveries in the BPA-nephro-vascular system paradigm, all this provides a sufficient basis to place kidney patients in the critical spotlight. The systematic review has determined relatively high BPA plasma values in patients undergoing hemodialysis, which could be a potentiating element for its worsening. In this way, the need to modify the materials used in specific treatments to reduce exposure to this endocrine disruptor is determined, thus avoiding some patients' possible deterioration. Similarly, the high values of urinary BPA in various publications related to occupational exposure show the need to improve personal protective equipment and working conditions in specific sectors related to the manufacture or recycling of plastics, since concentrations should not be detected urinary levels high enough to reach the micromolar or even nanomolar range. Although we have indeed normalized the existence of endocrine disruptors in the general

population's urine, which is a worrying fact, we should ensure that they are at the lowest possible threshold.

EFSA determines that the TDI is at four μg/kg BW/day, which is justified with experimental animal models. However, as in vitro experiments have shown, BPA can exert very different actions in murine and human cells, although with similar consequences, converging on the possibility of kidney damage. It remains to be determined whether it would be necessary to review the coherence of the calculations and extrapolations, taking into account the observed inter-species differences.

#### **8. Conclusions**


#### **Acknowledgements**

This work was supported in part by grants from Instituto de Salud Carlos III (PI15/02,139) -Fondo Europeo de Desarrollo Regional (FEDER)-.

R. Moreno- Gómez-Toledano is recipient of a research contract from CAM (B2017-BMD-3686).

*Critical Analysis of Human Exposure to Bisphenol A and Its Novel Implications on Renal… DOI: http://dx.doi.org/10.5772/intechopen.96309*

#### **Author details**

Rafael Moreno-Gómez-Toledano1 , María I. Arenas3 , Sandra Sánchez-Esteban<sup>2</sup> , Alberto Cook<sup>2</sup> , Marta Saura2 and Ricardo J. Bosch1 \*

1 Laboratory of Renal Physiology and Experimental Nephrology, University of Alcalá, Alcalá de Henares, SPAIN

2 Laboratory of Pathophysiology of the Vascular Wall, Department of System Biology/Physiology Unit, University of Alcalá, Alcalá de Henares, SPAIN

3 Department of Biomedicine and Biotechnology/Cell Biology Unit, University of Alcalá, Alcalá de Henares, SPAIN

\*Address all correspondence to: ricardoj.bosch@uah.es

© 2021 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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Section 2
