{"paper_id":"40c91a8e-0598-44f2-b54f-f4f39a52a6ae","body_text":"The phthalate esters are a class of water-insoluble, high-production-volume, synthetic organic chemicals\nused widely in a variety of industrial applications, including personal-care products (e.g., perfumes,\nlotions, cosmetics), paints, and mainly as plasticizers to confer flexibility and durability to polyvinyl\nchloride- (PVC-) based plastics and to make the plastic appropriate to different uses, including food, construction\nindustry, medical devices, and pharmaceuticals since about the 1930s [ 1 – 4 ]. However, these plasticizers are not chemically bound to the plastic products, but leak out from PVC items into the environment with time and use. As a consequence, they have been found everywhere in the environment and are universally considered ubiquitous environmental contaminants. Di-(2-ethylhexyl) phthalate (DEHP) is the most abundant phthalate in the environment and mono-(2-ethylhexyl) phthalate (MEHP) is its primary metabolite [ 1 – 4 ]. Other important phthalates production- and applicationwise are diethyl phthalate (DEP), dibutyl phthalate (DBP), di-iso- and di-n-butyl phthalate (DiBuP, DnBuP), butyl-benzyl phthalate (BBP), di-isononylphpthalate (DiNP) and\ndi-n-octyl phthalate (DnOP) [ 5 ]. Humans are exposed to phthalates for their whole lifetime, since intrauterine life [ 6 – 11 ].\nThe ability of these pollutants to affect human health is a major concern. In particular, evidence suggestive of harmful effects on the male reproductive system and related outcomes have gradually accumulated in recent years. In addition, there is wide demonstration that reproductive functions are altered by endocrine disrupting chemicals (EDCs), including phthalates. These chemicals have been found to interfere with the function of the endocrine system, which is responsible for growth, sexual development, and many other essential physiological functions in both genders.\nEDCs can act genomically, with agonistic or antagonistic effects on steroid receptors and may alter reproductive function and/or cause feminization by binding to oestrogen or androgen receptors. However, EDCs can also act by nongenomic mechanisms, altering steroid synthesis [ 12 ,  13 ].\nThe definition of endocrine disruption is today extended to broader endocrine regulations, and includes activation of metabolic sensors, such as a subset of nuclear hormone receptor superfamily members called\nperoxisome proliferator-activated receptors (PPARs).\nTo this regard, a large group of industrial and pharmaceutical chemicals, including phthalates, are\nknown for their ability to provoke peroxisome proliferation, thus increasing both\nthe size and number of peroxisomes [ 14 ]. Peroxisomes\nare essential organelles of eukaryotic origin, ubiquitously distributed in\ncells and organisms, which perform various metabolic functions (peroxide-derived\nrespiration, beta oxidation of fatty acids, cholesterol metabolism, etc.)\nwithin the cell [ 15 ].\nMany of the adaptive consequences for exposure to\nthese pollutants are mediated by PPARs, members of the nuclear hormone receptor\n(NRs) superfamily of ligand-activated transcription factors. They are activated\nby binding of natural ligands, such as polyunsaturated fatty acids or by\nsynthetic ligands. Three subtypes of PPARs (alpha, beta, and gamma) have been\nidentified in different tissues, encoded by separate genes [ 16 ].\nSeveral studies in recent years have revealed their importance in both normal physiology and in the pathology of various tissues [ 17 ,  18 ]. In particular, human and animal studies have demonstrated that PPARs are important in placental development [ 19 ], while they are believed to play an essential role in the adverse effects elicited by EDC [ 20 ].\nThe aim of this review is to explore how much evidence exists linking phthalate exposure, PPARs activation, and eventual actions of PPARs as mediators of environmental toxic substances for reproductive function in both genders.\n\nGlobally, more than 18 billion pounds of phthalates are used each year and well above two million tons of DEHP alone are produced annually worldwide [ 21 ]. Given their\nhigh production volume, common use, and widespread environmental contamination,\nhumans are exposed to these compounds through ingestion, inhalation, and dermal\nexposures on a daily basis as testified by detection of phthalates in serum, seminal fluid, amniotic fluid, breast\nmilk, and saliva [ 5 ,  9 ,  22 – 24 ]. These studies have provided evidence on the\nrelatively high variation of phthalate exposure from day to day within individuals as well as between ethnic groups, geographic areas, and ages. In particular, general population can be exposed to DEHP to a much higher extent than previously believed and an exposure of children, twice as high as the exposure of adults with respect to their body weight, has been observed [ 23 – 26 ].\nIn particular, higher DEHP exposure has been documented in neonatal intensive-care-unit infants,\nbecause of multiple medical device-related DEHP exposure [ 27 ].\nIn addition, Blount et al. [ 28 ] found that\nwomen of reproductive age had significantly higher urinary levels of MBP (a\nreproductive and developmental toxicant in rodents) than other age/gender\ngroups. However, in spite of the alarming wide environmental diffusion and use,\nstudies in human populations suggesting an association between phthalate exposure\nand adverse reproductive health outcomes are limited yet.\nTo this regard, chronic occupational exposure to high levels of phthalates is associated with\ndecreased rates of pregnancy and higher rates of miscarriage in female factory\nworkers [ 29 ,  30 ]. Correspondently,\nhigher urinary phthalate levels were observed to correlate with pregnancy complications\nsuch as anemia, toxemia, and pre-eclampsia in women living near a plastics manufacturer\n[ 31 ]. In addition, significantly high levels of phthalates were identified in girls with\nthelarche, suggesting an association between plasticizers with known estrogenic\nand antiandrogenic activity and the cause of premature breast development in a\nhuman female population [ 32 ].\nIn utero exposure to phthalates has been shown to be\nsignificantly associated with a shorter pregnancy duration [ 7 ,  8 ] and it has been hypothesized that phthalates may play a role in inducing and/or potentiating an intrauterine inflammatory response, a well established risk factor for prematurity [ 33 ]. Moreover,\nan association between phthalate exposure and endometriosis has been shown,\nsuggesting a potential role for phthalate esters in the pathogenesis of this\ncommon cause of female infertility [ 34 ,  35 ]. More\nspecifically to the male reproductive system, phthalate exposure seems to be\ntightly correlated to the impairment of androgen activity. For example,\nphthalate monoesters levels in breast milk resulted to be correlated with hormone\nlevels in healthy boys, which were indicative of lower androgen activity and\nreduced Leydig cell function [ 36 ], and\nprofessional long-term exposure to phthalates has been reported to be\nassociated with altered semen quality [ 37 ,  38 ] and decreased serum-free testosterone [ 39 ].\nIn addition, impaired testicular descent and decreased anogenital distance (AGD), the most sensitive marker of antiandrogen action in toxicological studies and a sensitive measure of prenatal antiandrogen\nexposure have been reported in boys whose mothers had elevated prenatal phthalate exposure [ 43 ]. All together, these findings suggest an impairment of sex hormone balance by prenatal\nand postnatal phthalate exposure but, although suggestive of the potentially\ndangerous effects of phthalate exposure on human health, they are not\nconclusive yet, and more epidemiologic data are needed in human populations along\nwith a better mechanistic understanding of the phthalates activities. Although the\npossible mechanism of action by phthalates remains, to date, largely obscure, the\nuse of animal models have enormously contributed to characterize the\nreproductive toxicity profiles of phthalates and to highlight the mechanisms possibly involved.\n\nMale and female reproductive tract development is a dynamic process, requiring the production and the fine regulatory activity of sex steroid hormones: androgens, estrogens, and the progestagens [ 40 ]. Steroidal sex hormones regulate foetal developmental processes such as differentiation\nand sex determination. The major sites of synthesis of the sex steroids are\ncorpus luteum for progestagens, testis for androgens, and ovaries for estrogens.\nThe biosynthesis of sex steroids is catalyzed by a series of enzymes that form the steroidogenic pathway [ 41 ]. This pathway causes the conversion of pregnenolone (cholesterol derivative key\nsteroidogenic intermediate common to all classes of steroid hormones) to\nprogesterone, the precursor for the testosterone that is formed in testis by\nLeydig cells through two ways: (1) Δ4-biosynthesis leads to progesterone,\n17- α -hydroxyprogesterone, and androstenedione; (2) the Δ5-biosynthesis leads to 17- α -hydroxypregnenolone, dehydroepiandrosterone, and Δ5-androstendiol [ 41 ].\nAndrogens themselves can then be transformed to estrogens. The extent to which this biotransformation takes place depends on the expression of the various enzymes in specific tissues. The enzyme complex\n19-hydroxylase-aromatase, which catalyzes the conversion of androgens to\nestrogens, plays a major role in this biotransformation [ 42 ].\nThe development of mammalian foetus into a male requires the \nproduction and action of steroid hormones, notably androgens \nand antimullerian hormone after testis formation, in contrast \nto the female development, a process largely hormone-independent \n[ 43 ].\nMoreover, the mature reproductive\nfunction is under the regulation of the hypothalamus-pituitary-gonadal (HPG)\naxis. The limbic system of the brain releases specific neurotransmitters or\nneuropeptides that stimulate the hypothalamus to produce gonadotropin-releasing\nhormone (GnRH) which stimulates the pituitary gland to release specific hormones (gonadotrophins) that are transported via the blood stream to hormone-synthesizing tissues [ 44 ]. In the case\nof mammals, the gonadotrophins from the pituitary gland are luteinizing\nhormone (LH) and follicle-stimulating hormone (FSH). Under the influence of\nthese substances, sex steroids, that is, estrogens and androgens, are released\ninto the blood circulation from the ovaries and the testis, respectively.\nNegative feedback from the concentration of these gonadal steroids in the blood\ncan lower or block the release of GnRH from the hypothalamus and of\ngonadotrophins at the pituitary level, thus modulating HPG axis [ 44 ].\nKeeping this in mind, it might be expected that\nany environmental, hormonally active chemicals capable of perturbing the\nadequate production and action of sex hormones or the balance between estrogens\nand androgens during foetal life have the potential to interfere with one or\nmore critical aspects of reproductive function ( Figure 1 ).\n\nChronic exposure of laboratory animals to phthalates has been reported to lead to severe adverse effects, including foetal death, carcinogenesis, teratogenesis, and hepatotoxicity [ 45 – 47 ]. In particular, a wide range of developmental and reproductive toxicities in mammals are induced by phthalates. Phthalates\ncan directly affect fetal and neonatal testis differentiation, inducing male rat reproductive tract malformations, as well as testicular changes remarkably similar to testicular dysgenesis\nsyndrome (TDS) in humans [ 48 – 52 ].\nTesticular dysgenesis, or abnormal testicular development, after in utero\nphthalate exposure has been shown to be associated with abnormal function of\nboth Sertoli and Leydig cells and abnormal sex organs development [ 52 ,  53 ].\nSertoli cells play a critical role in foetal testis development regulating the dynamic process of movement, organization, differentiation of all the cell types within the testis [ 54 ]. As a consequence, the abnormal function of Sertoli cells associated with phthalate\nexposure [ 52 ,  53 ] might alter\nthe differentiation signals normally implicated in tissue morphogenesis, thus\nleading to many of the histological and functional anomalies observed in TDS ( Figure 1 ).\nLeydig cells, the principal providers of steroid hormones in the testis, are also targeted by phthalates. To this regard, the highly conserved role of testosterone and dihydrotestosterone (DHT), in driving\nmale reproductive tract development (masculinization) is well known. As a\nconsequence, in rodents the whole period of male genital tract differentiation\nis particularly susceptible to the effects of antiandrogens, as demonstrated by in utero exposure to flutamide, (a well-known androgen receptor antagonist) and phthalates both inducing abnormalities\nof androgen-regulated sexual differentiation [ 49 ]. In\naddition, the administration of synthetic estrogens, such as diethylstilboestrol (DES), to pregnant women and rodents causes reproductive tract abnormalities in the offspring, including cryptorchidism, [ 55 ] as well as a dose-dependent reduction in the number of Sertoli cells critically involved in\nspermatogenesis [ 56 ]. The ability of estrogens to reduce androgen levels or expression of androgen receptor is relevant [ 57 ]. These\nresults suggest that abnormal intrauterine hormone levels with decreased\nandrogen production/action or increased estrogens levels may play a role in\ndetermining adverse effects on reproductive health. Correspondently, critical to\nthe induction of phthalate testicular toxicity is the considerable reduction in\nfetal and postnatal testosterone levels observed after in utero exposure to\nphthalates at the critical window for the androgen-dependent reproductive tract\ndevelopment [ 49 ,  52 ,  53 ,  58 ]. In particular, the exposure to DEHP decreases testosterone to levels similar to\nthose normally found in females leading to incomplete masculinization and hypospadias\nand cryptorchidism [ 58 ]. Thus, several phthalate\nesters have been shown to carry out “antiandrogenic” activity through a mechanism that is distinct from androgen-receptor antagonism, that is, targeting the\nLeydig cells testosterone biosynthesis machinery. In addition, genes directly\nassociated with testosterone biosynthesis are uniformly downregulated by\nphthalate exposure in the fetal testis [ 59 ]. These steroidogenic genes include those involved in cholesterol handling, such as scavenger receptor class\nB type 1 (SR-B1) implicated in the selective cholesterol esters uptake from\nhigh density lipoproteins, steroidogenic acute regulatory protein (StAR), that\nmediates cholesterol transport across the mitochondrial membrane, the rate\nlimiting enzyme in testosterone biosynthesis, that is, cholesterol side-chain\ncleavage enzyme (P450 scc), that converts cholesterol into pregnenolone, 3 β -hydroxysteroid dehydrogenase (3  β HSD), and CYP17 α  [ 59 ,  60 ]. In addition, phthalates alter the expression of genes encoding sex steroid metabolizing enzymes in the\ngonads and peripheral organs such as the liver. Among these, 5 α -reductase, that converts testosterone to DHT,\nwas upregulated by DEHP in the prepubertal rat testis [ 61 ]. Aside from the interference\nwith steroid synthesis and metabolism, the induction of cryptorchidism by\nphthalates is mediated by the alternative mechanism acting at the initial\nhormone-independent phase of testicular descent. Phthalates have indeed been\nshown to alter the expression of insulinlike hormone 3 (Insl3) in fetal Leydig\ncells [ 62 ], which plays a role in\nguiding the testis during its first phase of transabdominal descent.\nIn postnatal exposure, a strong species difference in the phthalate responsiveness is evident, with some species\n(Syrian hamsters, e.g.,) more resistant to phthalate toxicity possibly as a\nconsequence of an inefficient metabolic transformation of diesters to\nmonoesters [ 63 ]. Younger\nanimals result, in general, more sensitive than adult ones [ 64 ]. For\nexample, Grey observed a decrease in seminiferous tubule diameter in testis and\naccessory sex organs (seminal vesicle and prostate) weight after phthalate\nexposure in 4-week-old, but not in 15-week-old rats [ 64 ]. These\neffects were associated with the induction of apoptosis in germ cells, likely as\na consequence of an increased generation of oxidative stress and concomitant\nalteration of antioxidant defences by phthalate [ 65 ].\nCorrespondently, the FSH signalling pathway for Sertoli cell proliferation and\ndifferentiation resulted to be impaired after phthalate exposure [ 66 ,  67 ].\nAlso in postnatal and adult rats phthalates\naffected steroid hormone synthesis and metabolism, as indicated by decreased\ntestosterone serum levels in male rats acutely exposed to some active\nphthalates and by a decreased testosterone secretion by cultured Leydig cells\ntreated with MEHP [ 68 ]. However, contrasting\nresults were observed by Akingbemi et al. [ 69 ] and Eagon et al. [ 70 ] in male rat chronically exposed to environmentally relevant low levels of DEHP. Increased LH\nand testosterone serum levels together with an increased serum estrogen likely\ndue to impaired Leydig cell steroidogenesis and compensatory Leydig cell\nproliferation were observed. The modulation by phthalate of many estrogen\nmetabolizing enzymes seems to be very complex, since it has been reported both\na downregulation [ 71 ,  72 ] and an upregulation [ 73 ] of the aromatase gene after\nphthalate exposure, depending on the cell type analyzed.\nOverall, the data presented here demonstrated\nthat certain phthalates like other environmental chemicals are capable of\ndisrupting male reproductive tract organogenesis and function when administered\nto laboratory animals during pregnancy and/or postnatal life, producing types\nof malformations and histological changes causing infertility remarkably similar\nto those observed in human TDS. One mechanism responsible for this effects may\nbe the ability to disrupt the endocrine balance, that is, androgen/estrogen\nactivities, essential for reproductive system development and homeostasis,\nacting as environmental antiandrogen compounds [ 74 ]. Although\nthis raises concern towards other factors such as lifestyle that might have\ninfluenced human fertility [ 75 ].\n\nThe identification of phthalates as environmental chemicals belonging to the family of peroxisome proliferators (PP) has shed new insight into the potential molecular mechanism of phthalate action in the\nreproductive system of mammals. The pleiotropic effects induced by PP including\nphthalates in the rodent liver are mediated by the activation of PPARs, ligand-activated\ntranscription factors belonging to the nuclear receptor superfamily, which also\nincludes the steroid and thyroid hormone receptors [ 76 ]. Thus far,\nthree PPAR isoforms ( α ,  β , or  δ , and  γ ), encoded by separate genes, have been identified in various tissues, with PPAR α \npredominantly expressed in the liver, PPAR γ  in adipose tissue, and PPAR β  in a wider range of tissue [ 16 ]. Upon\nactivation by their lipophilic ligands, PPARs regulate gene transcription by\nbinding to PPAR response elements (PPRE) within the promoter of target genes as\nheterodimers with retinoic X receptors (RXR) [ 16 ,  77 ]. PPARs can also repress gene expression in a DNA-binding-dependent way through the recruitment of corepressors to unliganded PPARs as well as in a\nDNA-binding-independent manner by interfering with other nuclear signalling\npathways via protein-protein interaction (leading to formation of inactive\ncomplexes) or via competition for limiting amounts of the heterodimerization\npartner RXR or coactivators [ 78 ]. Fatty acids\nand eicosanoids have been identified as natural ligands for PPARs. More potent\nsynthetic PPAR ligands include the fibrate and thiazolidinedione drugs,\nclinically used as hypolipidemic and antidiabetic agents, respectively. Since\nthe discovery of PPARs in 1990 [ 17 ], several\nfunctions have been attributed to these receptors. PPARs play critical\nphysiological roles regulating lipid and glucose homeostasis, cellular\ndifferentiation, proliferation, and the inflammatory/immune response, with\nsubsequent clinically relevant implication in several diseases including dyslipidemia,\ndiabetes, cancer, atherosclerosis. PPAR α  has been\ndemonstrated to play a role in regulating lipid catabolism, whereas PPAR γ  controls\nadipocyte differentiation and lipid storage [ 16 ,  77 ]. Although PPAR β  is less well understood, it might be a mediator in the control of brain lipid metabolism, fatty acid-induced adipogenesis, and atherogenic inflammation [ 77 ]. Given the extensive crosstalk between PPARs and other transcription factors and signalling events regulating energy balance, differentiation and other\nsignificant physiological processes in many tissues, the involvement of\nenvironmental chemicals in the PPAR system may potentially result in\npathophysiologically relevant consequences for human health.\nThe role of PPAR α  in PP-induced hepatic proliferative responses was established by the development of PPAR α -deficient mice by Lee et al. [ 79 ]. In contrast to wild-type control animals, PPAR α  homozygous-deficient mice do not exhibit hepatic peroxisomal proliferation in response to treatment with PP. Aside from modest changes in lipid profile and weight, PPAR α -deficient mice are otherwise\nphenotypically normal [ 80 ]. Thus, the major hepatic effects of PP, including hepatocarcinogenic effects, are mediated by PPAR α -dependent\ngene transcription and signalling events. The response to PP seems to be\nspecies-specific, with rats and mice being quite sensitive to them and humans,\nguinea pigs, and other species being refractory [ 80 ]. Remarkably,\nthe hepatotoxic effects of PP are lost in humans due to the lower level of PPAR α  expression\nin human liver than in rodent one [ 81 ] and to\nspecies-specific responsiveness of PPAR α  [ 82 ].\nBefore focusing on the potential involvement of PPARs in the reproductive effects of phthalate, it would be useful to consider PPAR expression pattern in the reproductive system, since the potential\nPPAR-mediated effects of phthalates depend on tissue distribution of the PPAR\nisoforms and the PPAR-responsive genes in each tissue. All PPAR isoforms are\nexpressed in the central nervous system and in reproductive tissues, such as\ngonads (testis and ovary), uterus, prostate, mammary gland, pituitary gland [ 83 ]. In the\ntestis, both somatic and germ cells express PPAR isoforms: PPAR α  and  β \nare expressed in Leydig cells and cells of seminiferous tubule (Sertoli cells and germ cells) [ 60 ,  84 ], while PPAR γ  seems to be only detectable in Sertoli cells,\nalthough weak PPAR γ  expression in germ cells has recently been reported [ 85 ]. All PPAR isoforms have been detected in the ovary [ 84 ]. PPAR γ  is the predominant isoform expressed in the granulosa cells and preovulatory follicles, but its expression falls after the\nLH surge [ 86 ]. In addition, PPAR γ  is less strongly expressed in the techal cells and in corpus luteum where it increases after ovulation [ 86 ]. However, in\nthe absence of fertilization or embryo implantation, PPAR γ  expression decreases as a result of corpus\nluteum regression [ 87 ]. Finally, PPAR γ  is\nexpressed in uterine tissue, blastocyst and, together with PPAR α  and  β , in gestational tissues [ 88 ,  89 ].\nThe physiological role of PPARs in the reproductive tissues is not completely understood but while, on one hand, PPAR α -null mice remain viable and fertile [ 79 ], on the\nother hand, PPAR β  deletion impairs fertility [ 90 ] and PPAR γ -null mutation is even embryonically lethal [ 91 ]. Indeed,\nrecent findings suggested putative important roles for PPARs in reproductive system:\nthe ability of PPARs to regulate energy balance may represent a potential\nmolecular link between reproductive function and glucose and lipid metabolism.\nIt has been shown that PPAR α , whose expression is upregulated by FSH in cultured seminiferous tubules [ 92 ], may affect spermatozoa fertility by promoting lipid storage mobilization and modifying\nphospholipid composition. PPAR β  seems to play an important role in embryo implantation as showed by its strong upregulation during the decidualization process and the appearance of placental\nmalformations in PPAR β -null mice [ 90 ]. Finally,\nseveral lines of evidence suggest that PPAR γ  is critically involved in follicular\ndevelopment, ovulation, maintenance of corpus luteum during pregnancy, and\nmaturation and function of placenta [ 83 ].\n\nThe involvement of phthalate-PPAR interactions in the reproductive \nbiology alteration derives from recent findings demonstrating that \nphthalates are able to activate PPAR α  and PPAR γ  isoforms. Metabolic conversion of diesters to\nthe hydrolytic monoesters seems to be essential to obtain PPAR activation and\ntoxicological effects [ 93 ]. Indeed,\nhepatic peroxisomal proliferation and the associated hepatocarcinogenic\nresponse induced in rodents by DEPH are mediated by its bioactive metabolite\nMEHP [ 94 ], which is able to activate both human and rodent PPAR α  and PPAR γ  in in vitro transactivation assay [ 95 ]. In addition\nto MEHP, other structurally diverse phthalate monoesters, most notably\nmonobenzyl phthalate (mBzP), the primary metabolite of butyl benzyl phthalate\n(BBP), and mono- sec -butyl phthalate (MBuP) are capable of activating both human PPAR isoforms and target genes [ 93 ,  96 ] with\npotential implication for human health as these reproductive toxicants have\nbeen detected in human urine samples at exceptionally higher levels than MEHP\nitself [ 28 ]. However, it has been recently found that the diesters DEHP and BBP themselves were able to activate PPAR α  and PPAR γ  to some\nextent, although it was likely attributable to low level of esterases activity\nin the cell model used [ 96 ]. Interestingly, analyses of structure-activity relationship have found that PP in general are amphipathic carboxilates thus resembling natural PPAR ligands\nsuch as long-chain saturated and unsaturated fatty acids [ 97 ]. The\ncarboxyl moiety of monoesters is critical for ligand activity: for example,\nsome DEHP metabolites, such as MEHP and 2-ethylhexanoic acid, are more potent\nPPAR activators than 2-etylhexanol metabolite [ 98 ]. The rank\norder for phthalate activation of mouse and human PPAR α  and PPAR γ  agrees with the relative ability of phthalate esters to induce the classical PPAR responses, that are liver peroxisomal\nproliferation in rodents for PPAR α  and adipocyte differentiation for PPAR γ  [ 93 ,  99 ]. Indeed, it\nhas been found that esters with long and branch-side chain are more potent PPAR\nactivators than those containing short-chains or straight-chains. \nAs regards\nPPAR β , only phthalate monoesters with longer and branch-side chains can\nactivate this isoform but at a concentration higher than that \nrequired for activation of PPAR α  and PPAR γ  [ 100 ]. Importantly, human PPARs are less sensitive to phthalate monoesters than the corresponding mouse receptors [ 93 ]. Since the\nactivation of PPAR assessed by transactivation assay might result from indirect\nevents, such as endogenous production of a metabolite from the test compound or\nrelease of endogenous ligand, these compounds had to be tested further for\ndirect binding to the PPARs. Although activation of PPARs by some phthalates\nmay occur indirectly through release of endogenous lipid activators (fatty\nacids) from carrier proteins, notably fatty acid binding protein (FABP) or\nthrough a yet unidentified intermediate factor [ 101 ], recent\nfindings reported that some relevant monoester phthalates are able of directly\nbinding PPAR α  and PPAR γ  receptors [ 96 ]. Consistent\nwith their ability to activate PPARs in transactivation assay, BBP and DBP\nweakly interact with both isoforms.\nAlthough in most cases there has been found a correlation between PPAR activation by phthalate monoesters and reproductive toxicity by the corresponding diesters, there exist also findings weakening the assumption of a general obligatory role for PPARs in mediating phthalate-induced reproductive effects. For example, while di-isononyl phthalate (DINP) is a weak reproductive toxicant [ 102 ], its monoester metabolite MINP is a moderately strong PPAR activator [ 100 ]. In addition, DBP is a strong reproductive toxicant through its proximal metabolite MBP [ 103 ] and induces hepatotoxicity in rodents via PPAR α  [ 104 ], although MBP only weakly activates PPARs in transactivation assay [ 93 ]. One possible interpretation of these discordant results may be the involvement of\nan indirect mechanism of PPAR activation mediated by an unknown endogenous\nmetabolite activator, not necessarily detectable by using transactivation\nassay.\nOnly a few studies in PPAR α -null mice directly determined the role of PPAR\nin phthalate-induced male developmental and reproductive toxicities. The study by\nPeters et al. [ 105 ] showed that\nprenatal exposure to DEHP caused developmental malformations in both wild-type\nand PPAR α  knockout mice, thus suggesting a PPAR α -independent\nmechanism. However, it is difficult to draw any conclusion about the role of\nPPAR α  in phthalate reproductive toxicity since the intrauterine administration of DEHP\noccurred before the critical period of reproductive tract differentiation.\nAnother important animal study demonstrated that intrauterine DEHP-treated PPAR α -deficient mice, predominantly normal at earlier time point, developed delayed testicular, renal and developmental toxicities,\nbut not liver toxicity, compared to wild types [ 104 ], thus first\nconfirming the early observation by Lee et al. about the PPAR α \ndependence of liver response and, more importantly, indicating that DEHP may\ninduce reproductive toxicity through both PPAR α -dependent and -independent mechanism. Another\nstudy found that the administration of DEHP resulted in milder testis lesions\nand higher testosterone levels in PPAR α -null mice than in wild-type mice [ 106 ]. In contrast, the PPAR α -independent reproductive toxicity observed by Ward et\nal. may conceivably be mediated by other PPAR isoforms, such as PPAR β  and PPAR γ , or by a nonreceptor-mediated organ-specific mechanism. Unfortunately, till now no studies have been performed in PPAR β -null mice, and the toxicological impacts of phthalates that activate PPAR γ  are\nunknown. Determining a role for PPAR γ  in phthalate-induced reproductive toxicity requires testis-specific-knockout mice as PPAR γ  deletion results in the death of the embryo [ 91 ]. Notably, both PPAR α  and PPAR γ  are responsive to DEHP in vitro and\nare translocated to the nucleus in primary Sertoli cells after incubation of\nthese cells with phthalate esters [ 107 ,  108 ]. Given the\nkey role played by Sertoli cells in driving testis morphogenesis, it may be therefore hypothesized that the impairment of this cell type by MEHP contributed to the observed testicular toxicity.\nThe potential of PPARs to mediate the endocrine\ndisruption activity by phthalates is also suggested from the finding that a few\ngenes involved in steroid biosynthesis and metabolism are directly regulated by\nPPARs. MEHP activates both PPAR α  and PPAR γ  in\ncultured rat granulosa cells which cause a complete inhibition of aromatase\ngene expression [ 109 – 111 ]. In addition, the estradiol metabolizing enzyme 17 β -HSD IV has been shown to be induced\nby MEHP in the liver and granulosa cells through a PPAR α -dependent\nmechanism [ 112 ]. Therefore, both decreased estradiol synthesis and increased estradiol metabolism contribute to suppressed serum estradiol levels observed after DEHP in vivo exposure and to the subsequent female reproductive toxicity [ 71 ,  72 ,  113 ]. Finally, the induction by DEHP of FABP expression in the liver via PPAR α  [ 114 ] and in granulosa cells via both PPAR α  and PPAR γ  [ 115 ] may play important role in the mechanism of phthalate effect on steroid hormones since FABP functions as an intracellular gateway for PPAR agonists [ 116 ] and as a donor of potential fatty acid ligands of PPARs [ 101 ].\nTaking into account the specific tissue distribution and the physiological roles of PPAR isoforms, one could speculate upon some phthalate effects in mammals. It is known that cells exposed to PP\nundergo oxidative stress possibly due to PPAR α -mediated activation of metabolizing enzymes in\nthe liver and associated with the hepatic toxicity of DEHP [ 117 ]. Genes\ninvolved in oxidative stress response have been shown to be upregulated in the\nliver by DEHP exposure [ 118 ]. In addition,\nthe induction of xenobiotic metabolizing enzymes by PPAR α  after DEHP exposure could increase the\nsusceptibility to other environmental toxicants requiring metabolic activation [ 118 ]. PPAR γ  is a prototypic adipocyte differentiation regulator [ 119 ] and\nactivation of PPAR γ  by phthalates in other tissue and subsequent alteration of differentiation\npathways may be implicated in phthalate teratogenic effects. In addition, PPAR γ  may be part of the LH-induced luteinization in the ovary since its activation causes aromatase downregulation, this event\nbeing essential for the postovulatory phenotype [ 120 ]. The\nactivation of PPAR γ  by phthalates in the preovulatory follicle prevented the estradiol increase\nnecessary for stimulating the ovulatory surge of LH and prematurely induces\nfollicle differentiation to a postovulatory phenotype [ 113 ].\n\nThe above-mentioned epidemiological evidence suggesting adverse consequences for female reproductive function [ 30 ,  31 ] stimulated more\nin depth studies in animal models on the issue. Besides causing developmental\ntoxicity, including high incidence of foetus death and malformations and\nreduced foetal body weight, DEHP administration to pregnant rodents decreased\nembryo implantation and increased resorptions [ 121 ,  122 ]. These effects were mimicked by other phthalate esters thus representing both male and\nfemale reproductive toxicants in rodents [ 123 ].\nThe administration of phthalate esters, including DEHP and its metabolite MEHP, to adult female rats caused an increase in the estrous cycle length and dysovulation, associated with polycystic ovaries, and\ndecreased serum levels of estradiol [ 71 ]. These\nfunctional changes were associated with morphological alteration of the\npreovulatory follicle, the site of estradiol production, where granulosa cells\nwere smaller in DEHP-treated mice than in control rats, and incapable of\nmounting an ovulatory surge of LH. Regarding the molecular mechanism by which\nDEHP/MEHP suppressed estradiol production in the granulosa cells, it has been\nfound that MEHP inhibits FSH-stimulated cAMP accumulation and progesterone\nproduction in granulosa cells [ 124 ]. When the\nprogesterone precursor pregnenolone is added to granulosa cell cultures treated\nwith MEHP, the inhibition of progesterone production is reversed [ 125 ]. However\nMEHP did not decrease the expression of P450 scc [ 126 ], the major\nregulatory site of progesterone production by cAMP which converts cholesterol to\npregnenolone [ 127 ]. In addition\nto reducing progesterone production at a site prior to pregnenolone, MEHP also\nreduces estradiol production by affecting aromatase gene expression, the\nrate-limiting enzyme that converts testosterone to estradiol. Aromatase is\nstimulated by FSH-mediated pathways and techal androgens. Androgens are the\nsubstrates for aromatization to estradiol in granulosa cells [ 128 ]. Thus, MEHP\nis able to decrease estradiol production independent of its effect on FSH–cAMP and\ndecreases aromatase activity without acting as a direct enzyme inhibitor [ 72 ].\nFurthermore, the induction by both DEHP and DBP of the estradiol metabolizing\nenzyme 17 β -HSD IV in the liver and granulosa cells [ 112 ,  129 ] contributes to explain the suppressed serum estradiol levels after DEHP exposure and the\nsignificant increase in serum levels of estrone, the primary metabolite of\nestradiol, observed in DBP-treated rats [ 71 ].\nOverall, these findings underline once again that phthalate toxicant effects on female reproductive system is attributable to an interference with the complex and tightly regulated machinery involved in\nsteroid synthesis and metabolism. Notably, the pathways leading to production\nof ovarian hormones are similar in rodent models and humans, and using the\nrodent model to determine the mechanism of action of MEHP will aid in\nunderstanding how exposure to this chemical may affect ovarian function in\nwomen.\n\nPhthalates are environmental contaminants with significant human exposures. These chemicals may act as EDCs and alter reproductive function and/or cause feminization raising concern about the potential health hazards posed by such exposures. The adverse effects of phthalates have been chiefly studied in animal models, while their potential toxicity to humans together with the possible involvement of PPARs in mediating these effects on the reproductive health has to be more properly evaluated. Pre- and/or perinatal periods appear to be critical windows of exposure, because of their high sensitivity to hormonal dysregulation by EDCs. Thus, the acquisition of more detailed data on human exposure during these time periods is essential. It has been proposed that impairment of reproductive development and function in both genders by phthalates relates to abnormal steroid biosynthesis and metabolism and seems to be at least in part mediated by the activation of the PPAR signalling pathway. Molecular basis for the adverse health effects proposed to be associated with human phthalate exposure have to be elucidated. Finally, analysis of the effects of phthalate exposures on gonadotropin and steroid hormone levels should form part of overall risk assessment in human populations.","source_license":"CC-BY-4.0","license_restricted":false}