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PPAR Res. 2008; 2008: 359267. Published online 2007 December 6. doi: 10.1155/2008/359267. | PMCID: PMC2225463 |
Peroxisome Proliferator-Activated Receptors as Mediators of Phthalate-Induced Effects in the Male and Female Reproductive Tract: Epidemiological and Experimental Evidence Giuseppe Latini,1, 2* Egeria Scoditti,2, 3 Alberto Verrotti,4 Claudio De Felice,5 and Marika Massaro2, 3 1Division of Neonatology, Perrino Hospital, 72100 Brindisi, Italy 2Clinical Physiology Institute (IFC-CNR), National Research Council of Italy, Lecce Section, 72100 Brindisi, Italy 3Laboratory of General Physiology, Department of Biological and Environmental Sciences and Technology, University of Lecce, 73100 Lecce, Italy 4Department of Medicine, Division of Pediatrics, University of Chieti, 66100 Chieti, Italy 5Neonatal Intensive Care Unit, Azienda Ospedaliera-Universitaria Senese, 53100 Siena, Italy Received May 28, 2007; Revised September 12, 2007; Accepted September 25, 2007. This is an open access article distributed under the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. There is growing evidence that male as well as female reproductive function has been declining in human and wildlife populations over the last 40 years. Several factors such as lifestyle or environmental xenobiotics other than genetic factors may play a role in determining adverse effects on reproductive health. Among the environmental xenobiotics phthalates, a family of man-made pollutants are suspected to interfere with the function of the endocrine system and therefore to be endocrine disruptors. The definition of endocrine disruption is today extended to broader endocrine regulations, and includes activation of metabolic sensors, such as the peroxisome proliferator-activated receptors (PPARs). Toxicological studies have shown that phthalates can activate a subset of PPARs. Here, we analyze the epidemiological and experimental evidence linking phthalate exposure to both PPAR activation and adverse effects on male and female reproductive health. The phthalate esters are a class of water-insoluble, high-production-volume, synthetic organic chemicals
used widely in a variety of industrial applications, including personal-care products (e.g., perfumes,
lotions, cosmetics), paints, and mainly as plasticizers to confer flexibility and durability to polyvinyl
chloride- (PVC-) based plastics and to make the plastic appropriate to different uses, including food, construction
industry, 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
di-n-octyl phthalate (DnOP) [ 5]. Humans are exposed to phthalates for their whole lifetime, since intrauterine life [ 6– 11]. The 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. EDCs 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]. The 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
peroxisome proliferator-activated receptors (PPARs). To this regard, a large group of industrial and pharmaceutical chemicals, including phthalates, are
known for their ability to provoke peroxisome proliferation, thus increasing both
the size and number of peroxisomes [ 14]. Peroxisomes
are essential organelles of eukaryotic origin, ubiquitously distributed in
cells and organisms, which perform various metabolic functions (peroxide-derived
respiration, beta oxidation of fatty acids, cholesterol metabolism, etc.)
within the cell [ 15]. Many of the adaptive consequences for exposure to
these pollutants are mediated by PPARs, members of the nuclear hormone receptor
(NRs) superfamily of ligand-activated transcription factors. They are activated
by binding of natural ligands, such as polyunsaturated fatty acids or by
synthetic ligands. Three subtypes of PPARs (alpha, beta, and gamma) have been
identified in different tissues, encoded by separate genes [ 16]. Several 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]. The 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. 2. ENVIRONMENTAL DISSEMINATION AND EPIDEMIOLOGICAL EVIDENCE OF PHTHALATE REPRODUCTIVE TOXICITY Globally, 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
high production volume, common use, and widespread environmental contamination,
humans are exposed to these compounds through ingestion, inhalation, and dermal
exposures on a daily basis as testified by detection of phthalates in serum, seminal fluid, amniotic fluid, breast
milk, and saliva [ 5, 9, 22– 24]. These studies have provided evidence on the
relatively 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]. In particular, higher DEHP exposure has been documented in neonatal intensive-care-unit infants,
because of multiple medical device-related DEHP exposure [ 27]. In addition, Blount et al. [ 28] found that
women of reproductive age had significantly higher urinary levels of MBP (a
reproductive and developmental toxicant in rodents) than other age/gender
groups. However, in spite of the alarming wide environmental diffusion and use,
studies in human populations suggesting an association between phthalate exposure
and adverse reproductive health outcomes are limited yet. To this regard, chronic occupational exposure to high levels of phthalates is associated with
decreased rates of pregnancy and higher rates of miscarriage in female factory
workers [ 29, 30]. Correspondently,
higher urinary phthalate levels were observed to correlate with pregnancy complications
such as anemia, toxemia, and pre-eclampsia in women living near a plastics manufacturer
[ 31]. In addition, significantly high levels of phthalates were identified in girls with
thelarche, suggesting an association between plasticizers with known estrogenic
and antiandrogenic activity and the cause of premature breast development in a
human female population [ 32]. In utero exposure to phthalates has been shown to be
significantly 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,
an association between phthalate exposure and endometriosis has been shown,
suggesting a potential role for phthalate esters in the pathogenesis of this
common cause of female infertility [ 34, 35]. More
specifically to the male reproductive system, phthalate exposure seems to be
tightly correlated to the impairment of androgen activity. For example,
phthalate monoesters levels in breast milk resulted to be correlated with hormone
levels in healthy boys, which were indicative of lower androgen activity and
reduced Leydig cell function [ 36], and
professional long-term exposure to phthalates has been reported to be
associated with altered semen quality [ 37, 38] and decreased serum-free testosterone [ 39]. In 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
exposure 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
and postnatal phthalate exposure but, although suggestive of the potentially
dangerous effects of phthalate exposure on human health, they are not
conclusive yet, and more epidemiologic data are needed in human populations along
with a better mechanistic understanding of the phthalates activities. Although the
possible mechanism of action by phthalates remains, to date, largely obscure, the
use of animal models have enormously contributed to characterize the
reproductive toxicity profiles of phthalates and to highlight the mechanisms possibly involved. 3. MALE AND FEMALE REPRODUCTIVE TRACT DEVELOPMENT: POSSIBLE
INTERFERENCE SITE BY PHTHALATES Male 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
and sex determination. The major sites of synthesis of the sex steroids are
corpus luteum for progestagens, testis for androgens, and ovaries for estrogens. The 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
steroidogenic intermediate common to all classes of steroid hormones) to
progesterone, the precursor for the testosterone that is formed in testis by
Leydig cells through two ways: (1) Δ4-biosynthesis leads to progesterone,
17- α-hydroxyprogesterone, and androstenedione; (2) the Δ5-biosynthesis leads to 17- α-hydroxypregnenolone, dehydroepiandrosterone, and Δ5-androstendiol [ 41]. Androgens 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
19-hydroxylase-aromatase, which catalyzes the conversion of androgens to
estrogens, plays a major role in this biotransformation [ 42]. The development of mammalian foetus into a male requires the
production and action of steroid hormones, notably androgens
and antimullerian hormone after testis formation, in contrast
to the female development, a process largely hormone-independent
[ 43].
Moreover, the mature reproductive
function is under the regulation of the hypothalamus-pituitary-gonadal (HPG)
axis. The limbic system of the brain releases specific neurotransmitters or
neuropeptides that stimulate the hypothalamus to produce gonadotropin-releasing
hormone (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
of mammals, the gonadotrophins from the pituitary gland are luteinizing
hormone (LH) and follicle-stimulating hormone (FSH). Under the influence of
these substances, sex steroids, that is, estrogens and androgens, are released
into the blood circulation from the ovaries and the testis, respectively.
Negative feedback from the concentration of these gonadal steroids in the blood
can lower or block the release of GnRH from the hypothalamus and of
gonadotrophins at the pituitary level, thus modulating HPG axis [ 44]. Keeping this in mind, it might be expected that
any environmental, hormonally active chemicals capable of perturbing the
adequate production and action of sex hormones or the balance between estrogens
and androgens during foetal life have the potential to interfere with one or
more critical aspects of reproductive function (). 4. PRE- AND POSTNATAL DEVELOPMENTAL AND REPRODUCTIVE TOXICITY BY PHTHALATES Chronic 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
can directly affect fetal and neonatal testis differentiation, inducing male rat reproductive tract malformations, as well as testicular changes remarkably similar to testicular dysgenesis
syndrome (TDS) in humans [ 48– 52]. Testicular dysgenesis, or abnormal testicular development, after in utero
phthalate exposure has been shown to be associated with abnormal function of
both Sertoli and Leydig cells and abnormal sex organs development [ 52, 53]. Sertoli 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
exposure [ 52, 53] might alter
the differentiation signals normally implicated in tissue morphogenesis, thus
leading to many of the histological and functional anomalies observed in TDS (). Leydig 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
male reproductive tract development (masculinization) is well known. As a
consequence, in rodents the whole period of male genital tract differentiation
is 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
of androgen-regulated sexual differentiation [ 49]. In
addition, 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
spermatogenesis [ 56]. The ability of estrogens to reduce androgen levels or expression of androgen receptor is relevant [ 57]. These
results suggest that abnormal intrauterine hormone levels with decreased
androgen production/action or increased estrogens levels may play a role in
determining adverse effects on reproductive health. Correspondently, critical to
the induction of phthalate testicular toxicity is the considerable reduction in
fetal and postnatal testosterone levels observed after in utero exposure to
phthalates at the critical window for the androgen-dependent reproductive tract
development [ 49, 52, 53, 58]. In particular, the exposure to DEHP decreases testosterone to levels similar to
those normally found in females leading to incomplete masculinization and hypospadias
and cryptorchidism [ 58]. Thus, several phthalate
esters have been shown to carry out “antiandrogenic” activity through a mechanism that is distinct from androgen-receptor antagonism, that is, targeting the
Leydig cells testosterone biosynthesis machinery. In addition, genes directly
associated with testosterone biosynthesis are uniformly downregulated by
phthalate exposure in the fetal testis [ 59]. These steroidogenic genes include those involved in cholesterol handling, such as scavenger receptor class
B type 1 (SR-B1) implicated in the selective cholesterol esters uptake from
high density lipoproteins, steroidogenic acute regulatory protein (StAR), that
mediates cholesterol transport across the mitochondrial membrane, the rate
limiting enzyme in testosterone biosynthesis, that is, cholesterol side-chain
cleavage enzyme (P450  scc), that converts cholesterol into pregnenolone, 3 β-hydroxysteroid dehydrogenase (3 βHSD), and CYP17 α [ 59, 60]. In addition, phthalates alter the expression of genes encoding sex steroid metabolizing enzymes in the
gonads and peripheral organs such as the liver. Among these, 5 α-reductase, that converts testosterone to DHT,
was upregulated by DEHP in the prepubertal rat testis [ 61]. Aside from the interference
with steroid synthesis and metabolism, the induction of cryptorchidism by
phthalates is mediated by the alternative mechanism acting at the initial
hormone-independent phase of testicular descent. Phthalates have indeed been
shown to alter the expression of insulinlike hormone 3 (Insl3) in fetal Leydig
cells [ 62], which plays a role in
guiding the testis during its first phase of transabdominal descent. In postnatal exposure, a strong species difference in the phthalate responsiveness is evident, with some species
(Syrian hamsters, e.g.,) more resistant to phthalate toxicity possibly as a
consequence of an inefficient metabolic transformation of diesters to
monoesters [ 63]. Younger
animals result, in general, more sensitive than adult ones [ 64]. For
example, Grey observed a decrease in seminiferous tubule diameter in testis and
accessory sex organs (seminal vesicle and prostate) weight after phthalate
exposure in 4-week-old, but not in 15-week-old rats [ 64]. These
effects were associated with the induction of apoptosis in germ cells, likely as
a consequence of an increased generation of oxidative stress and concomitant
alteration of antioxidant defences by phthalate [ 65].
Correspondently, the FSH signalling pathway for Sertoli cell proliferation and
differentiation resulted to be impaired after phthalate exposure [ 66, 67]. Also in postnatal and adult rats phthalates
affected steroid hormone synthesis and metabolism, as indicated by decreased
testosterone serum levels in male rats acutely exposed to some active
phthalates and by a decreased testosterone secretion by cultured Leydig cells
treated with MEHP [ 68]. However, contrasting
results 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
and testosterone serum levels together with an increased serum estrogen likely
due to impaired Leydig cell steroidogenesis and compensatory Leydig cell
proliferation were observed. The modulation by phthalate of many estrogen
metabolizing enzymes seems to be very complex, since it has been reported both
a downregulation [ 71, 72] and an upregulation [ 73] of the aromatase gene after
phthalate exposure, depending on the cell type analyzed. Overall, the data presented here demonstrated
that certain phthalates like other environmental chemicals are capable of
disrupting male reproductive tract organogenesis and function when administered
to laboratory animals during pregnancy and/or postnatal life, producing types
of malformations and histological changes causing infertility remarkably similar
to those observed in human TDS. One mechanism responsible for this effects may
be the ability to disrupt the endocrine balance, that is, androgen/estrogen
activities, essential for reproductive system development and homeostasis,
acting as environmental antiandrogen compounds [ 74]. Although
this raises concern towards other factors such as lifestyle that might have
influenced human fertility [ 75]. 5. THE PPAR SYSTEM AT THE CROSSROADS BETWEEN METABOLISM AND
REPRODUCTION The 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
reproductive system of mammals. The pleiotropic effects induced by PP including
phthalates in the rodent liver are mediated by the activation of PPARs, ligand-activated
transcription factors belonging to the nuclear receptor superfamily, which also
includes the steroid and thyroid hormone receptors [ 76]. Thus far,
three PPAR isoforms ( α, β, or δ, and γ), encoded by separate genes, have been identified in various tissues, with PPAR α
predominantly expressed in the liver, PPAR γ in adipose tissue, and PPAR β in a wider range of tissue [ 16]. Upon
activation by their lipophilic ligands, PPARs regulate gene transcription by
binding to PPAR response elements (PPRE) within the promoter of target genes as
heterodimers 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
DNA-binding-independent manner by interfering with other nuclear signalling
pathways via protein-protein interaction (leading to formation of inactive
complexes) or via competition for limiting amounts of the heterodimerization
partner RXR or coactivators [ 78]. Fatty acids
and eicosanoids have been identified as natural ligands for PPARs. More potent
synthetic PPAR ligands include the fibrate and thiazolidinedione drugs,
clinically used as hypolipidemic and antidiabetic agents, respectively. Since
the discovery of PPARs in 1990 [ 17], several
functions have been attributed to these receptors. PPARs play critical
physiological roles regulating lipid and glucose homeostasis, cellular
differentiation, proliferation, and the inflammatory/immune response, with
subsequent clinically relevant implication in several diseases including dyslipidemia,
diabetes, cancer, atherosclerosis. PPAR α has been
demonstrated to play a role in regulating lipid catabolism, whereas PPAR γ controls
adipocyte 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
significant physiological processes in many tissues, the involvement of
environmental chemicals in the PPAR system may potentially result in
pathophysiologically relevant consequences for human health. The 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
phenotypically normal [ 80]. Thus, the major hepatic effects of PP, including hepatocarcinogenic effects, are mediated by PPAR α-dependent
gene transcription and signalling events. The response to PP seems to be
species-specific, with rats and mice being quite sensitive to them and humans,
guinea pigs, and other species being refractory [ 80]. Remarkably,
the hepatotoxic effects of PP are lost in humans due to the lower level of PPAR α expression
in human liver than in rodent one [ 81] and to
species-specific responsiveness of PPAR α [ 82]. Before 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
PPAR-mediated effects of phthalates depend on tissue distribution of the PPAR
isoforms and the PPAR-responsive genes in each tissue. All PPAR isoforms are
expressed in the central nervous system and in reproductive tissues, such as
gonads (testis and ovary), uterus, prostate, mammary gland, pituitary gland [ 83]. In the
testis, both somatic and germ cells express PPAR isoforms: PPAR α and β
are 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,
although 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
LH 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
the absence of fertilization or embryo implantation, PPAR γ expression decreases as a result of corpus
luteum regression [ 87]. Finally, PPAR γ is
expressed in uterine tissue, blastocyst and, together with PPAR α and β, in gestational tissues [ 88, 89]. The 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
other hand, PPAR β deletion impairs fertility [ 90] and PPAR γ-null mutation is even embryonically lethal [ 91]. Indeed,
recent findings suggested putative important roles for PPARs in reproductive system:
the ability of PPARs to regulate energy balance may represent a potential
molecular link between reproductive function and glucose and lipid metabolism.
It 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
phospholipid 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
malformations in PPAR β-null mice [ 90]. Finally,
several lines of evidence suggest that PPAR γ is critically involved in follicular
development, ovulation, maintenance of corpus luteum during pregnancy, and
maturation and function of placenta [ 83]. 6. MECHANISM OF PHTHALATE ESTER REPRODUCTIVE TOXICITY:
POTENTIAL ROLE OF PPARS The involvement of phthalate-PPAR interactions in the reproductive
biology alteration derives from recent findings demonstrating that
phthalates are able to activate PPAR  and PPAR  isoforms. Metabolic conversion of diesters to
the hydrolytic monoesters seems to be essential to obtain PPAR activation and
toxicological effects [ 93]. Indeed,
hepatic peroxisomal proliferation and the associated hepatocarcinogenic
response induced in rodents by DEPH are mediated by its bioactive metabolite
MEHP [ 94], which is able to activate both human and rodent PPAR α and PPAR γ in in vitro transactivation assay [ 95]. In addition
to MEHP, other structurally diverse phthalate monoesters, most notably
monobenzyl phthalate (mBzP), the primary metabolite of butyl benzyl phthalate
(BBP), and mono- sec-butyl phthalate (MBuP) are capable of activating both human PPAR isoforms and target genes [ 93, 96] with
potential implication for human health as these reproductive toxicants have
been detected in human urine samples at exceptionally higher levels than MEHP
itself [ 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
extent, although it was likely attributable to low level of esterases activity
in 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
such as long-chain saturated and unsaturated fatty acids [ 97]. The
carboxyl moiety of monoesters is critical for ligand activity: for example,
some DEHP metabolites, such as MEHP and 2-ethylhexanoic acid, are more potent
PPAR activators than 2-etylhexanol metabolite [ 98]. The rank
order 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
proliferation in rodents for PPAR α and adipocyte differentiation for PPAR γ [ 93, 99]. Indeed, it
has been found that esters with long and branch-side chain are more potent PPAR
activators than those containing short-chains or straight-chains.
As regards
PPAR β, only phthalate monoesters with longer and branch-side chains can
activate this isoform but at a concentration higher than that
required 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
activation of PPAR assessed by transactivation assay might result from indirect
events, such as endogenous production of a metabolite from the test compound or
release of endogenous ligand, these compounds had to be tested further for
direct binding to the PPARs. Although activation of PPARs by some phthalates
may occur indirectly through release of endogenous lipid activators (fatty
acids) from carrier proteins, notably fatty acid binding protein (FABP) or
through a yet unidentified intermediate factor [ 101], recent
findings reported that some relevant monoester phthalates are able of directly
binding PPAR α and PPAR γ receptors [ 96]. Consistent
with their ability to activate PPARs in transactivation assay, BBP and DBP
weakly interact with both isoforms. Although 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
an indirect mechanism of PPAR activation mediated by an unknown endogenous
metabolite activator, not necessarily detectable by using transactivation
assay. Only a few studies in PPAR α-null mice directly determined the role of PPAR
in phthalate-induced male developmental and reproductive toxicities. The study by
Peters et al. [ 105] showed that
prenatal exposure to DEHP caused developmental malformations in both wild-type
and PPAR α knockout mice, thus suggesting a PPAR α-independent
mechanism. However, it is difficult to draw any conclusion about the role of
PPAR α in phthalate reproductive toxicity since the intrauterine administration of DEHP
occurred before the critical period of reproductive tract differentiation.
Another 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,
but not liver toxicity, compared to wild types [ 104], thus first
confirming the early observation by Lee et al. about the PPAR α
dependence of liver response and, more importantly, indicating that DEHP may
induce reproductive toxicity through both PPAR α-dependent and -independent mechanism. Another
study found that the administration of DEHP resulted in milder testis lesions
and 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
al. 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
unknown. 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
are translocated to the nucleus in primary Sertoli cells after incubation of
these cells with phthalate esters [ 107, 108]. Given the
key 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. The potential of PPARs to mediate the endocrine
disruption activity by phthalates is also suggested from the finding that a few
genes involved in steroid biosynthesis and metabolism are directly regulated by
PPARs. MEHP activates both PPAR α and PPAR γ in
cultured rat granulosa cells which cause a complete inhibition of aromatase
gene expression [ 109– 111]. In addition, the estradiol metabolizing enzyme 17 β-HSD IV has been shown to be induced
by MEHP in the liver and granulosa cells through a PPAR α-dependent
mechanism [ 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]. Taking 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
undergo oxidative stress possibly due to PPAR α-mediated activation of metabolizing enzymes in
the liver and associated with the hepatic toxicity of DEHP [ 117]. Genes
involved in oxidative stress response have been shown to be upregulated in the
liver by DEHP exposure [ 118]. In addition,
the induction of xenobiotic metabolizing enzymes by PPAR α after DEHP exposure could increase the
susceptibility to other environmental toxicants requiring metabolic activation [ 118]. PPAR γ is a prototypic adipocyte differentiation regulator [ 119] and
activation of PPAR γ by phthalates in other tissue and subsequent alteration of differentiation
pathways 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
being essential for the postovulatory phenotype [ 120]. The
activation of PPAR γ by phthalates in the preovulatory follicle prevented the estradiol increase
necessary for stimulating the ovulatory surge of LH and prematurely induces
follicle differentiation to a postovulatory phenotype [ 113]. 7. DEVELOPMENTAL AND REPRODUCTIVE TOXICITY OF PHTHALATES IN FEMALE
ANIMAL MODELS The above-mentioned epidemiological evidence suggesting adverse consequences for female reproductive function [ 30, 31] stimulated more
in depth studies in animal models on the issue. Besides causing developmental
toxicity, including high incidence of foetus death and malformations and
reduced foetal body weight, DEHP administration to pregnant rodents decreased
embryo implantation and increased resorptions [ 121, 122]. These effects were mimicked by other phthalate esters thus representing both male and
female reproductive toxicants in rodents [ 123]. The 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
decreased serum levels of estradiol [ 71]. These
functional changes were associated with morphological alteration of the
preovulatory follicle, the site of estradiol production, where granulosa cells
were smaller in DEHP-treated mice than in control rats, and incapable of
mounting an ovulatory surge of LH. Regarding the molecular mechanism by which
DEHP/MEHP suppressed estradiol production in the granulosa cells, it has been
found that MEHP inhibits FSH-stimulated cAMP accumulation and progesterone
production in granulosa cells [ 124]. When the
progesterone precursor pregnenolone is added to granulosa cell cultures treated
with MEHP, the inhibition of progesterone production is reversed [ 125]. However
MEHP did not decrease the expression of P450 scc [ 126], the major
regulatory site of progesterone production by cAMP which converts cholesterol to
pregnenolone [ 127]. In addition
to reducing progesterone production at a site prior to pregnenolone, MEHP also
reduces estradiol production by affecting aromatase gene expression, the
rate-limiting enzyme that converts testosterone to estradiol. Aromatase is
stimulated by FSH-mediated pathways and techal androgens. Androgens are the
substrates for aromatization to estradiol in granulosa cells [ 128]. Thus, MEHP
is able to decrease estradiol production independent of its effect on FSH–cAMP and
decreases aromatase activity without acting as a direct enzyme inhibitor [ 72].
Furthermore, the induction by both DEHP and DBP of the estradiol metabolizing
enzyme 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
significant increase in serum levels of estrone, the primary metabolite of
estradiol, observed in DBP-treated rats [ 71]. Overall, 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
steroid synthesis and metabolism. Notably, the pathways leading to production
of ovarian hormones are similar in rodent models and humans, and using the
rodent model to determine the mechanism of action of MEHP will aid in
understanding how exposure to this chemical may affect ovarian function in
women. Phthalates 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.
| Table 1Structures and related name of the most common phthalate monoesters.
Diesters of o-phthalic acid are quickly metabolized in vivo to their active metabolites, the monesters. The length and structure of the side chain are important for toxicity. |
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AND POSTNATAL DEVELOPMENTAL AND REPRODUCTIVE TOXICITY BY PHTHALATES
