1. Synthetic Phenols, Parabens, and Phthalates
The human population is continually exposed to a wide range of chemical compounds present in the environment, including phenols, parabens, and phthalates (PPPs), which can affect human health.
The most studied phenol is bisphenol A (BPA, 2,2-bis [4-hydroxyphenyl]propane), a synthetic chemical used in consumer products such as the lining of canned food and drinks, the packaging of baby formula and baby bottles, and dental implants [1]. BPA can migrate from cans into food, from plastic bottles into water, or from dental fillings into saliva, leading to an oral exposure. In addition, BPA can penetrate the skin through thermal paper in sales receipts and by inhalation via industry. Despite its ban from plastic baby bottles in 2011 in Europe (
Since the banning of BPA in many consumer products [1], substitutes have been proposed by manufacturers over the last decade, including bisphenol S (BPS, 4,4′-sulfonyldiphenol) and bisphenol F (BPF, 4,4′-dihydroxydiphenylmethane). BPS and BPF structures are similar to BPA, suggesting that they can act as endocrine disruptors as well. They have already been shown to cause obesity effects/metabolic effects in animal models [4], but further experiments are needed to assess their impact on organ function and also during pregnancy on the phenotype of the offspring, as these compounds have only been used for a short time.
The phenol family also includes triclosan (TCS, 5-chloro-2-(2,4-dichlorophenoxy) phenol), a synthetic antimicrobial agent [5]. TCS is used in several personal care products, including toothpaste, antibacterial soaps, dishwashing liquids, deodorant soaps, cosmetic and antiseptic products, and deodorants. TCS is also used in kitchen utensils, toys, bedding, clothes, fabrics, and trash bags. The main route of human exposure is dermal exposure [6], but exposure can also occur through the consumption of water and/or food products [7]. TCS is metabolized to glucuronide and sulfate conjugates, and is mainly excreted in the urine in humans [8]. Several investigators have reported that TCS is an EDC that affects the immune system, and cardiovascular and reproductive functions [7,9,10,11,12]. Thus, the percentage of TCS in certain products has been regulated by the EU, and it is also banned in shaving products [13]. Since TCS is used in many products, it is important to know the effects of direct exposure to this molecule on the phenotype as well as the potential intergenerational effects.
Parabens are considered as endocrine disruptors [14] and used to prevent the growth of microorganisms in personal care products such as lotions, deodorants, hair care products, shaving products, pharmaceuticals, textiles, and clothes, but also in food additives such as methylparaben (MeP, food additive E218), ethylparaben (EtP, E214), and, to a lesser extent, propylparaben (PrP, E216), heptylparaben (HeP), butylparaben (n-BuP), and isobutylparaben (i-BuP) [14]. The estimated total intake of parabens from foods achieves 307–940 ng/kg bodyweight [15]. The most common parabens investigated and detected in biological fluids are MeP, EtP, n-PrP, and n-BuP, whereas i-BuP, isopropylparaben (i-PrP), and benzylparaben (BzP) are less investigated. Moreover, the urine concentrations of parabens are higher in women compared to men. Therefore, it is essential to test the effects of exposure to all of these molecules on the phenotypes at key stages of development and during pregnancy so to evaluate the effects on the offspring.
Phthalates are the main plasticizers used in the industry to improve the extensibility, elasticity, and workability of polymers. Phthalates include low-molecular-weight di-n-butyl phthalate (DBP), benzyl butylphthalate (BBP), di-(2-ethylhexyl) phthalate (DEHP), di-methyl phthalate (DMP), and diethyl phthalate (DEP), or high-molecular-weight di-isononyl phthalate (DiNP) and di-isodecyl phthalate (DiDP). More than 25 phthalates are used in many consumer products, such as toys, clothing, inks, rubbers, adhesives, paints, household materials, pesticides, cosmetics, personal care products, pharmaceutical products, etc. [16,17,18]. They are present in our environment and are able to reach the food and beverage chain by migration during processing, storing, transportation, and preparation. Even if ingestion is considered the main source of human exposure, dermal or nasal routes are not negligible. Several phthalates have been classified by the European Commission as having endocrine-disrupting or reprotoxic properties (category 1B). In the EU, several phthalates are now banned from use in cosmetics, including BBP, DEHP, DBP, and bis(2-methoxyethyl) phthalate (DMEP) (EC/1223/2009). The use of several compounds is also regulated in material intended to come into contact with food (DEHP, DBP, and BBP, EC/2007/19) as well as in toys and childcare articles (DEHP, DBP, BBP, DiNP, DiDP, and di-n-octyl phthalate (DnOP), French law 2006-1361) [16,18,19]. In Europe, the REACH list limits the level of exposure to DEHP metabolites, including (mono(2-ethyl-5-oxohexyl) phthalate (MEOHP), MEHP, and mono(2-ethyl-5-hydroxyhexyl) phthalate (MEHHP), within 0.02 mg/kg bw/d. Noteworthy, the EFSA has set the tolerable daily intakes (in mg/kg of bw) to 0.05 for DEHP, 0.01 for DBP, 0.15 for DiNP and DiDP, and 0.5 for BBP [20,21]. The half-life of phthalate diesters in blood plasma or urine is less than 24 h, and these compounds are mainly excreted in the urine as conjugated monoesters [22], but some of them, such as DEHP, undergo a secondary metabolism before urinary excretion. Consequently, urine is the more relevant matrix for the biomonitoring of human phthalate exposure [18], even if all the metabolites of a parent molecule cannot be assayed in this fluid for technical reasons. Exposure to phthalates is raising a great concern regarding their impact on human health. For several years, accumulating evidence has suggested that phthalates and their metabolites are suspected to act as EDCs through their interaction with different signaling pathways. Phthalate exposure has been correlated to several health outcomes, among them adverse reproductive disorders in women and men, neurological diseases, cancer, metabolic disorders, asthma, allergies, and toxicity in both the liver and kidney [16], but the intergenerational effects should be investigated further with doses mimicking the human exposure.
In this review, we will focus on PPPs, i.e., their metabolism, the toxicokinetic models for adapting human exposure doses to animal models, and the transplacental transfer of PPPs. The effects of PPP exposure throughout critical periods such as preconception, gestation, and lactation on offspring health, including the placental function, neonatal phenotypes, metabolic outcomes, cardiometabolic disorders, and gonadal functions, will be presented in the context of the developmental origins of health and disease (DOHaD) concept [23]. These data will be based on the literature from epidemiological data, associating human maternal urinary concentrations of these molecules and their metabolites with the health effects on the offspring, and with regard to the long-term postnatal cardiometabolic effects, they will be mainly presented through studies on animal models, since human cohorts currently rarely explore the effects beyond childhood.
2. Absorption, Distribution, Metabolism, and Excretion Processes of PPPs
PPPs are considered non-persistent chemicals because absorbed PPPs are rapidly metabolized and primarily excreted in the urine. In humans, ingested BPA and its analogs are rapidly and efficiently absorbed, and almost entirely excreted in the urine as a conjugate [24,25,26,27].
Following oral administration to human volunteers, TCS is rapidly absorbed, metabolized to conjugates, and eliminated in the urine within 4 days of exposure, with a terminal plasma half-life of 21 h [28]. Toxicokinetic studies performed in rats have shown that benzophenone-3 (BP-3) is rapidly absorbed after oral exposure and mainly excreted in the urine in conjugated forms [29]. BP-3 can also be transformed into three major metabolites, namely, benzophenone-1 (BP-1), benzophenone-8 (BP-8), and 2,3,4-trihydroxybenzophenone (THB), with BP-1 being the most frequently detected in rats [30,31]. Parabens are well absorbed following oral or dermal exposure and extensively metabolized through hydrolysis into para-hydroxybenzoic acid (PHBA), which is excreted in the urine mainly as glucuronide and sulfate conjugates. A fraction of absorbed parabens is also excreted in the urine in the form of glucuronide and sulfate conjugates [32]. Since PHBA is a non-specific metabolite of parabens, several biomonitoring studies have assessed the human exposure by measuring the urinary concentrations of conjugated parabens.
For phthalates, mono-n-butyl phthalate (MBP), mono-isobutyl phthalate (MiBP), and mono-benzyl phthalate (MBzP) are the main urinary metabolites of their respective parent short-chain phthalates, including di-n-butyl phthalate (DnBP), diisobutyl phthalate (DiBP), and BBP, respectively. The monoester metabolites of long-chain phthalates, such as DEHP and DiNP, MEHP, and MiNP, respectively, give rise to secondary oxidative metabolites (mono-2-ethyl-5-hydroxy-hexyl phthalate (MEHHP), mono-2-etyl-5oxohexyl phthalate (MEOHP), mono-2-ethyl-5-carboxypentyl phthalate (MECPP), and mono [2-(carboxymethyl)hexyl] phthalate (MCMHP)) for DEHP, and several complex isoforms of monoisononyl phthalate (MiNP) obtained by the formal condensation of one of the carboxy groups of phthalic acid with the hydroxy group of isononanol, which are the main xenobiotic metabolites excreted in human urine either as free or conjugated compounds [33].
Because PPPs are rapidly metabolized and excreted in the urine, several biomonitoring studies have assessed the human exposure by measuring the urinary concentrations of total (unconjugated plus conjugated) phenols and phthalate monoesters. Indeed, urine represents a relevant biological matrix for the biomonitoring of non-persistent chemical substances such as PPPs because their concentrations in urine are higher than in plasma concentrations, and because its collection is non-invasive. Generalized human exposure to PPPs is therefore reflected by the high frequency of detection of phenols and phthalate monoesters in the urine [2,34,35,36,37,38].
3. Toxicokinetic Models for Adapting Human Exposure Doses to Animal Models
Toxicokinetics (TKs) provides critical information for integrating hazard toxicity and exposure assessments to determine the potential risk of exposure to PPPs in humans. TKs is needed for determining the dose/concentration range that should be used in animal or in vitro testing to reproduce the internal exposure that can be expected from realistic human external exposure scenarios.
3.1. Estimation of PPP Daily Intake
PPP urine biomonitoring data have been widely used as biomarkers of exposure for epidemiological purposes [36,39]. Since urine biomonitoring data provide an integrative measure of exposure regardless of the source and route, the measurement of urinary concentrations of total PPP metabolites allows for the direct assessment of the daily PPP intake (DI). The calculation of the DI from urinary biomonitoring data relies on the urinary excretion factors (Fue) of the measured PPPs or metabolites related to the parent chemical ingested (Equation (1), [40]), as follows:
(1)
where UE is the molar urinary excretion of the measured PPPs or metabolites (µmol/L); UVnorm is the daily excreted urinary volume (L/kg bw/day), estimated at 0.028 L per kg bw per day for a 70-kg bodyweight [41], and which should be adapted to the targeted population (i.e., pregnant women and infants); FUE is the molar ratio between the amount of PPPs or metabolites excreted in the urine and the amount of parent PPPs taken up; MW is the molecular weight of the parent PPPs.For phenols, most biomonitoring studies rely solely on the determination of total phenol concentrations in the urine measured after enzymatic hydrolysis. While bisphenols and TCS are primarily eliminated in the urine as conjugates [25,28], the fraction of absorbed parabens excreted in the urine as glucuronides and sulfate conjugates ranges from about 6% for n-BuP to 17% for propylparaben (PrP) [42,43]. The selection of the appropriate biomarkers of exposure to long-chain phthalates such as DEHP and DiNP is more complex. Indeed, while approximately 70–85% of an oral dose of short-chain diester phthalates (DnBP, DiBP, BBP, and DEP) is excreted in the urine as hydrolytic monoester phthalates (MBP, MiBP, MBzP, and MEP) [44,45], only about 2 and 7% of an oral dose of DiNP and DEHP are excreted in the urine as the hydrolytic monoesters [40,46,47]. The secondary (oxidized monoester) phthalate metabolites of DiNP, such as OH-MiNP, oxo-MiNP, and cx-MiNP, and of DEHP, such as MECPP, MEOHP, MEHHP, and MCMHP, are the major metabolites excreted in human urine as free or conjugated compounds. Therefore, urinary concentrations of hydrolytic monoesters of long-chain relative to short-chain phthalates do not reflect the relative exposure to phthalates [48].
Although urine biomonitoring data provide essential information on the amount of compound that enters the body on a daily intake, regardless of the route of exposure, it is the concentration of the toxicologically active form of PPPs in the blood that determines the amount of chemicals that can reach the target tissues and receptors and exert effects.
For phenols, only the native (unconjugated) form is considered toxicologically active because it is capable of interacting with target tissues and receptors. Since phthalate diesters, such as DEHP, are cleaved to their hydrolytic monoesters in the mouth, stomach, intestines, or blood, plasma concentrations of phthalate diesters are expected to be very low [33]. It is the plasma concentrations of monoester phthalate, namely, MEHP, for DEHP that are considered relevant in terms of risk assessment [49].
3.2. Interspecies Extrapolation of PPP Dose
The steady-state plasma concentration (CSS) of the toxicologically active form of PPPs therefore represents the most relevant physiological variable for translating the results from animal studies to humans. CSS (ng/mL) is related to the dosing rate (mass per unit of time) by a key toxicokinetic parameter, namely, blood clearance (CL, volume per unit of time), in addition to the bioavailability (F, ranging from zero to one). This toxicokinetic parameter corresponds to the fraction of compound reaching the systemic circulation without change according to Equation (2), as follows:
(2)
Using Equation (2), the PPP dosage to be administered orally to reproduce the targeted CSS of the toxicologically active form of PPPs in an animal species can therefore be calculated from the species’ CL and F values.
A loading dose (LD) of PPPs can be applied to enable serum concentrations to rapidly reach the targeted CSS. This dose is determined from the steady-state volume of the distribution (Vss) and F using Equation (3), as follows:
(3)
By permitting the serial sampling of maternal and fetal blood after maternal or fetal administration, pregnant ewes provided a unique model for evaluating the disposition of several drugs during the prenatal period [50]. Among animal models, the rabbit represents an appropriate model for TK purposes. Like in humans, urinary elimination predominates in rabbits. This is unlike other species, such as rats and dogs, that have rather biliary excretion [51]. This interspecies difference is mainly observed for substances with a molecular weight between 300 and 600 Da, such as certain phthalate monoesters.
The main TK parameters of PPPs, i.e., CL, Vss, and F, can be determined from the plasma concentration–time profiles obtained following the intravenous and oral administrations of a mixture of EDCs. To achieve this, BPS and DEHP were administered by the oral route in a mixture of eight compounds in female rabbits to compute the loading and daily maintenance doses. The aim was to reproduce in rabbits the steady-state serum concentrations (CSS) of the two compounds found in pregnant women. Thus, the mean CSS of female rabbits measured after 17 weeks of exposure to a mixture of the eight compounds administered orally at doses deduced from the estimated TK parameters ranged from 0.77 to 1.21 ng/mL of the targeted CSS [52]. Toxicodynamic modeling in rabbits validated the interspecies extrapolation of human exposure rates to environmental contaminants.
4. Evaluation of Transplacental Transfer of PPPs Using Models
Understanding the materno-fetal transfer of PPPs across the placenta is crucial for assessing the prenatal exposure risks. To respond to this issue, the placental transfer of PPPs was investigated using the following two main models, which offer both advantages and disadvantages:
The most clinically relevant model to evaluate the placental transfer of compounds is the ex vivo human placental perfusion model at term [53]. This enables placental transfer to be assessed by reproducing maternal and fetal circulation in a few hours and taking samples from each compartment. This model offers several advantages, as the placental barrier maintains its structural integrity and the separate perfusion of the maternal and fetal sides reproduces their respective blood flows. However, this physiological approach does not take into account the nonplacental toxicokinetic factors that could contribute to the level of fetal exposure. These factors include the maternal and fetal metabolism, the dynamic structural conditions of the pregnancy, and the physiological changes throughout the pregnancy (the thickness of the trophoblast, uterine flow, and expression of transporters) [54].
An integrated pregnant sheep model that enables a direct administration and monitoring of xenobiotics over time in both fetal and maternal blood [55]. Because of important physiological similarities between the sheep and human placental functions, this model has contributed to significant advances in prenatal human medicine [56], despite the interspecies differences in the placental structure (the synepitheliochorial versus hemochorial placental structure) and transporters [56,57].
The ex vivo placental perfusion model had already been used to quantify the materno-fetal placental transfer of fifteen bisphenols [58]. Despite their chemical–structural similarities, these bisphenols differed greatly in the efficiency of placental transport. The placental transfer rates of bisphenol analogs such as 4,4′-Dihydroxybiphenyl, bisphenol AP (BPAP), bisphenol E (BPE), bisphenol F (BPF), 3-3BPA, bisphenol B (BPB), and BPA were similar to those of antipyrine, a molecule that passes the placental barrier by passive diffusion. This suggests that the exchange of these bisphenols across the placenta primarily involves passive diffusion and is limited only by the rate of the placental blood flow. In contrast, the placental transfer rates of bisphenol FL (BPFL) and BPS were very limited and intermediate for bisphenol BP (BPBP), bisphenol Z (BPZ), bisphenol C (BPC), bisphenol M (BPM), bisphenol P (BPP), and bisphenol AF (BPAF). This suggests a low diffusional permeability and/or that their passage might involve efflux transport. Moreover, although the glucuronide form of bisphenols could be detected in vivo in the fetal compartment, the placental transfer of glucuronide, evaluated for BPA and BPS, was almost non-existent. This indicates that, in the fetal compartment, these glucuronides come mainly from the fetoplacental metabolism, at least at the end of gestation [59,60]. A classical QSAR model based on molecular descriptors of BPs was developed to predict their materno-fetal transfer. But even if the physicochemical, topological, thermodynamic, and electronic parameters were able to influence the placental passage of this family of emerging BPs, the placental transfer efficiency of these 15 structurally related BPs could not be predicted solely from their physicochemical properties determined in silico [61]. This reinforces the indisputability of the physiological models.
Using the chronically catheterized fetal sheep model, toxicokinetic studies have been performed to determine the relative contributions of the placental transfer and the fetal and maternal metabolism pathways that control fetal exposure to BPA and BPS [55,60,62]. At the end of pregnancy, about 6% of the maternal dose of BPA reaches the fetal circulation, i.e., a dose related to the bodyweight equivalent to the maternal dose. This value was relative to the estimated 3.1% in humans using an ex vivo placental perfusion model [59]. Most of the BPA entering the fetal circulation was rapidly eliminated, mainly through the direct transfer of the BPA from the fetus to the mother (74%), and, to a lesser extent, through the fetal metabolism of BPA to BPA-glucuronide. In the fetus, entrapped BPA-glucuronide is eliminated by hydrolysis into the active form of BPA [62]. Therefore, the fetus has a much higher and sustained exposure to BPA metabolites.
It was striking that the fraction of the maternal BPS dose transferred from the mother to the fetus (0.40%) was about ten times lower than that of BPA [62]. This result is in agreement with the 10 times higher transfer efficiency of BPA from the mother to the fetus compared to that of BPS, as demonstrated in the ex vivo placental perfusion model [60]. Once in the fetal compartment, the percentage of the fetal dose of BPS entering the fetal blood eliminated by placental transfer was 26% compared to 74% for BPA. About half of the remaining BPS was metabolized to BPS-glucuronide by the fetus. The removal of BPS-glucuronide from the fetal compartment required its retro-conversion to bioactive BPS, like that of most bisphenol-glucuronides, due to the limited placental transfer. This toxicokinetic model predicted that, despite a lower materno-fetal passage of BPS compared to BPA, the higher persistence of BPS in the fetal compartment leads to expected BPS concentrations in fetal plasma of the same order of magnitude as that of BPA [63]. In agreement with these findings, the comparison of the fetal disposition of BPA, BPS, and BPF in the same pregnant ewe showed that the ratio of total feto-maternal concentrations of BPS was lower than those of BPF and BPA after a single subcutaneous administration of BPS or a mixture of BPA, BPS, and BPF [64].
To our knowledge, the placental transfer rate of TCS has not been studied using an ex vivo placental perfusion model or a pregnant sheep model. Concerning parabens, using a dual-recirculation placental perfusion model on term human placentae, Andersen et al. showed that the four parabens, MeP, EtP, PrP, and n-BuP, and their metabolite PHBA, just like BPA, are rapidly transported across the placental barrier. This implies potential fetal exposure [65].
Phthalate diesters are rapidly hydrolyzed to phthalate monoesters after maternal ingestion, inhalation, or dermal exposure in humans. In utero exposure to phthalates has been demonstrated in laboratory animals by measuring the concentrations of phthalate diesters or their metabolites in amniotic fluid [66], fetal plasma [66], and embryonic tissues [67]. In pregnant rats, the materno-fetal transfer of 14C-di(2-ethylhexyl)phthalate (DEHP) and 14C-diethyl phthalate (DEP) has been evaluated. This study showed that both phthalates and/or their metabolites were present in the placenta, amniotic fluid, and fetus [68]. Additionally, the pharmacokinetics of DBP in this animal model indicated that MBP was the major metabolite in the maternal and fetal plasma. This suggests that the active metabolite of DBP may cross the placenta in the late pregnancy. In amniotic fluid, the major metabolite was initially MBP, but by 24 h after administration, it was MBP-glucuronide [66]. Other studies confirmed that MBP was rapidly transferred to the embryonic tissues and that fetal plasma MBP-glucuronide concentrations were higher than those of the mother [67,69]. After the oral administration of DEHP to pregnant rats, unconjugated MEHP was the predominant metabolite measured in the amniotic fluid [70]. The transplacental transfer of MMP and MEHP, corresponding to the metabolites of DMP and DEHP, respectively, was evaluated in a dual-recirculation placental perfusion model of human placenta. This evaluation showed that MMP can cross the placenta by slow transfer, whereas no placental transfer was evidenced for MEHP. Moreover, MEHP, MBP, and MEP were found in fetal perfusate, corresponding to umbilical cord blood, whereas MMP, MEHHP, MBzP, and MiNP were undetectable. MEP, MBP, and MEHP were detected in placental tissue [71,72].
Altogether, these data demonstrated that the PPPs and their metabolites can cross the placenta, reach the fetal bloodstream, and contribute to the fetal exposome. Due to the short half-life of phthalate metabolites and the variation in the sensibility of the assay, the spot blood materno-fetal concentrations in term are difficult to interpret and cannot allow for the establishment of a relationship between the maternal and fetal concentrations, as shown for BPA [73,74]. Moreover, in the case of phthalates, it is not recommended to measure phthalates in the blood, since there are enzymes that can transform the parent into metabolites after blood collection, leading to false estimation.
5. Effects of PPPs on Human Placenta from Epidemiological Data and In Vitro Models and on Fetoplacental Growth from Animal Models
Only a few studies have looked at the association between PPPs and the markers of placental development. These are presented in the following paragraphs.
5.1. Epidemiological Data (Table 1)
Of the two studies investigating BPA exposure, one reported sex-specific associations (decreased female placental weight and increased placental weight in males, N = 232/220 male/female) [75], while the other, looking only at male births, showed no association with the placental weight nor with the placental-to-birthweight ratio, a marker of placental efficiency (N = 473) [36]. The single study looking at BPS did not report an association. This could be the result of the small sample size (N = 91) and the low frequency of detection of BPS in the urine (27%), thus limiting the power of the study [76]. Both studies on TCS reported negative associations with the placental weight; however, only one study reported it in male births [36], while the other study only reported it in female births [76]. Of the three studies investigating parabens (individual compounds or the molar sum), two studies reported a positive association with the placental weight in males at birth, either with n-BuP alone [76] or with the molar sum of the four parabens (including n-BuP) [36]. A third study evaluating paraben concentrations in placental tissues instead of maternal urine found opposite results (decreased placental weight with increased paraben placental concentrations) [77]. Of the two studies evaluating BP-3, only one study reported a positive association in male births [36], but the other study, perhaps because of its relatively small sample size (N = 91), did not report a positive association [76].
Regarding phthalates, a large cohort study (N = 2725 mother–child pairs from China) reported that, out of seven metabolites (MMP, MEP, MBP, MBzP, MEHP, MEOHP, and MEHHP), five (MMP, MBP, MEOHP, MEHHP, and MEHP) were associated with placental growth measurements, including the placental thickness, breadth, and length [78]. Most of these associations were positive, suggesting that prenatal exposure to phthalates can lead to a thicker and more circular placenta [78]. This cohort study reported stronger associations among male fetuses than in females. Moreover, the same cohort reported associations between all metabolites (N = 6: MMP, MEP, MBP, MEHP, MEOHP, and MEHHP), either individually or studied as the molar sum, and the placental weight [79]. The associations were also generally positive, particularly when exposure occurred late in pregnancy (the second and third trimesters). This cohort study also explored the effects of mixing, and although most phthalates were associated with placental outcomes when taken individually, no mixing effect was detected.
A few other studies with smaller sample sizes (ranging from 132 to 488) have also reported associations between certain phthalate metabolites and placental weight or the placental-to-birthweight ratio (PFR) [36,75,80]; however, the incriminated metabolites often differed from one study to another. Casas et al. reported a sex-specific association for MBzP that was positively associated with placental weight in boys and negatively associated in girls [75]. No association was reported for this metabolite in the other studies. Philippat et al. reported a negative association between cx-MiNP and MCOP and the placental weight or PFR in male fetuses [36], whereas this association was not reported in another study with a smaller sample size (N = 132, no stratification for child sex), which instead reported a decreased placental weight with prenatal exposure to MEP [80]. Interestingly, this study also reported a negative association between the preconception paternal exposure to DEHP and placental weight, and between the preconception maternal exposure to MEP and the birthweight-to-placental-weight ratio, providing preliminary results indicating the effect of the preconception exposure on placental development. Finally, a study evaluating phthalate metabolites in cord blood reported no association with the placental weight and volume [81].
Using the French mother–child cohort SEPAGES, 4 parabens, 2 bisphenols, 13 phthalate metabolites, and 2 non-phthalate plasticizer metabolites of DiNCH were measured in pools of repeated urine samples collected during the second and third trimesters of pregnancy. The results suggested negative associations between individual phthalate metabolites and the placental weight (MBzP and ΣDiNP), placental efficiency (MBzP, MBP, and ΣDiNP), and placental vascular resistance (MBzP, MBP, and ΣDiNP), indicating the adverse impacts of phthalate exposure on placental health [82]. Furthermore, using the same cohort, monoisobutyl phthalate (MiBP) and mono-n-butyl phthalate (MBP) were positively associated with most fetal growth parameters measured in the second trimester. Then, in the third trimester, MBP was further positively associated with the biparietal diameter and femur length [83].
Table 1Effects of phenols, parabens, and phthalates on the placenta using epidemiological data.
| Population | Placental Outcomes | Studied Compounds | Sample Type, Number, and Timing | Main Results | References | |||
|---|---|---|---|---|---|---|---|---|
| N = 2723 women–child pairs | Placental weight (calculated), PFR, chorionic plate area, disk eccentricity | MMP, MEP, MBP, ∑LMWP, MEHP, MEOHP, MEHHP, ∑DEHP | One spot urine sample per trimester | GM of three trimesters/T1/T2/T3 | [79] | |||
| N = 2725 (1399 boys and 1326 girls) mother–child pairs | Placental length, thickness, breadth and surface area | MMP, MEP, MBzP, MEHP, MEHHP, MEOHP | One maternal spot urine sample per trimester | T1/T2/T3 | GirlsPlacental breadth:MMP ↑*/↑/↑MBP ↑*/↑/↓LMWP ↑*/↑/↑Placental thickness:MBP ↑*/↑/↑*MEOHP ↑*/↑/↓ | [84] | ||
| N = 473 mother–infant (only boys) pairs | Placental weight, PFR | MCPP, MBP, MiBP, MBzP, MEP, cx-MiNP, MCOP, MEHP, MEHHP, MEOHP, MECPP, 2,4-DCP, 2,5-DCP, n-BuP, BP-3, BPA, EtP, MeP, PrP, TCS | One maternal spot urine sample collected between 23 and 29 gestational weeks | Penalized effects: | Unpenalized effects: | [36] | ||
| 132 mother–child pairs Subfertile population | Placental weight | MEP, MBP, MiBP, MBzP, MEHP, MEHHP, MEOHP, MECPP, MCPP, MCOP, cx-MiNP | Paternal and maternal spot urine samples collected: | [p-value] | [80] | |||
| N = 207 women–child pairs, recruited between October 2011 and September 2012 Chongqing (Southwest China) | Placental weight at birth | DMP, DEP, DMEP, DBP, DEEP, DiBP, DPP, BMPP, DBEP, DCHP, DnHP, BBP, DEHP, DnOP, DNP | Cord blood sample | No effect on placental weight or volume | [81] | |||
| N = 142 mother–newborn pairs | Placental weight at birth | MeP, EtP, PrP, n-BuP | Placenta sample | [p-value] | In girls:EtP ↓ Placental weight [0.02] ∑ Parabens ↓ Placental weight [0.03] | [77] | ||
| N = 657 mother–child pairs | Placental weight at birth | BPA, DEHP (∑ of MEHP, MEHHP, MEOHP, MECPP), MBzP, LMWP (∑ of MEP, MiBP, MBP | Two spot urine samples collected at 12 ± 1.7 and 32 ± 1.4 weeks of gestation | In boys: | In girls: | [75] | ||
| N = 130 mother–child pairs | Placental weight, fetal weight, head circumference, abdominal circumference, and femur length | 2,4-DCP, 2,5-DCP, BP-3, n-BuP, EtP, MeP, PrP, TCS, BPS, TRCB | One spot urine sample | Inverse associations were observed between average 2,4- and 2,5-DCP concentrations and birth weight z-scores in males | [76] | |||
| SEPAGES cohort | Placental weight, PFR, placental thickness, and placental vascular resistance | 4 parabens, 2 bisphenols, triclosan, benzophenone-3, 13 | Repeated urine samples collected during the second and third trimesters of pregnancy | Several phthalate metabolites were negatively associated with placental outcomes | [82] | |||
| SEPAGES cohort | Fetal biparietal diameter, femur length, head and abdominal circumferences measured by ultrasound, | 13 phthalates, and 1,2-cyclohexane dicarboxylic acid and diisononyl ester (DiNCH) metabolite | Repeated urine samples collected during the second and third trimesters of pregnancy | MiBP: ↑ biparietal diameter, and head and abdominal circumferences at T2 trimester | [83] | |||
BBP: benzyl butyl phthalate; BMPP: bis (4-methyl-2-pentyl) phthalate; n-BuP: n-butylparaben; BPA: bisphenol A; BP-3: benzophenone-3; DBP: di-n-butyl phthalate; DBEP: bis (2-nbutoxyethyl) phthalate; DCHP: dicyclohexyl phthalate; 2,4-DCP: 2,4-dichlorophenol; 2,5-DCP: 5-dichlorophenol; DEEP: bis (2-ethoxyethyl) phthalate; DEHP: di-(2-ethylhexyl) phthalate; DEP: diethyl phthalate; DnHP: di-n-hexyl phthalate; DiBP: diisobutyl phthalate; DiNCH: di-iso-nonyl-cyclohexane-1,2-dicarboxylate; DMP: di-methyl phthalate; DMEP: bis (2-methoxyethyl) phthalate; DnOP: di-n-octyl phthalate; DNP: dinonyl phthalate; DPP: di-amyl phthalate; EtP: ethylparaben; HMW: high-molecular-weight phthalates; LMW: low-molecular-weight phthalates; MBP: mono-n-butyl phthalate; MBzP: monobenzyl phthalate; cx-MiNP: mono-carboxy-iso-nonyl phthalate; MCOP: mono-carboxy-iso-octyl phthalate; MCPP: mono-3-carboxypropyl phthalate; MECPP: mono-2-ethyl-5-carboxypentyl terephthalate; MEHHP: mono-2-ethyl-5-hydroxyhexyl) phthalate; MEHP: mono-2-ethyl-hexyl) phthalate; MEOHP: mono(2-ethyl-5-oxohexyl) phthalate; MEP: mono-ethyl phthalate; MiBP: mono-iso-butyl phthalate; MMP: mono-methyl phthalate; MeP: methylparaben; PFR: placental-to-birthweight ratio; PrP: propylparaben; TCS: triclosan; TRCB: triclocarban. * indicates statistically significant results, ↓ indicates a decrease, ↑ indicates an increase.
5.2. Effects of PPPs on Human Placental Methylation
DNA methylation is the best characterized and most stable epigenetic modification, influencing the chromatin structure and the gene expression. This epigenetic mark typically involves the methylation of the fifth carbon position at a cytosine residue within a CpG dinucleotide (CpG), resulting in 5-methylcytosine (5mC). Recent studies have provided concrete evidence of a link between DNA methylation alterations and certain environmental exposures. Given the evidence that PPPs can cross the placenta, studies have explored the associations between PPP maternal exposure and the evolution of DNA methylation profiles in the placenta. The first studies focused on the locus of imprinted genes, IGF2 and H19 [85]. This locus displays an allele-specific expression (the expression of H19 from the maternal allele and of IGF2 from the paternal allele). The IGF2 gene and lncRNA H19 are together important for embryogenesis and fetoplacental development. The allele-specific methylation is involved in the allele-specific control of expression. Using placental samples collected at delivery from 196 healthy newborns, genomic DNA was used to analyze three differentially methylated regions, IGF2DMR0, IGFDMR2, and H19DMR, considered as imprinting the center region (H19 ICR). An increase in the level of methylation of the imprinting center region (H19 ICR) was associated with the sum of phthalate metabolites and low-molecular-weight metabolites. An increase in high-molecular-weight phthalate metabolites and DHEP exposure was associated with a deviation of the allele-specific expression of H19 (10%) only in male placenta. This suggests a potentially sexually dimorphic response. Conversely, the study showed that the sum of phthalate metabolites and high-molecular-weight metabolites was associated with a decrease in IGF2DMR0, without significant modification of the methylation level of IGF2DMR. In this study, the effects of the exposure to phenols were also investigated, but no significant association was found [85]. Quite different results were obtained by Zhao et al. [86] from the placental samples of 101 healthy newborns and 80 neonates with fetal growth retardation. The placental methylation of IGF2DMR was significantly inversely associated with MEHHP and MEOHP concentrations. Associations were much more evident in neonates with fetal growth restriction than in healthy neonates. Using the same mother–newborn cohort, measuring the phthalate exposure during the third trimester, the same authors demonstrated that the placental methylation of LINE-1 was negatively correlated with phthalate exposure levels. This outcome was significantly altered in cases of fetal growth restriction compared to healthy births. Because the LINE-1 repetitive element exhibits a wide distribution in the human genome, it has been frequently used as a surrogate marker of global methylation analysis. The modest hypomethylation of the LINE-1 repetitive element observed after exposure to phthalates could contribute to genome instability and increase the risk of chronic disease later in life [87].
A genome-wide DNA methylation analysis using the Illumina Infinium Human Methylation 450K or 850K Bead Chip was also performed. Although these differently methylated CpG sites constitute a small proportion of the total number of CpG sites in the genome (28 million), they represent a wide distribution of sites [88]. Grindler et al. used this tool to analyze genome-wide DNA methylation marks in placental villi in the first trimester from 16 women with high or low phthalate exposures [89]. They reported 2214 differentially methylated cytosines (DMCs) targeting 1460 unique genes. Taking into account the proximity of the DMCs, 282 differentially methylated regions (DMRs) were identified and found to be associated with 245 unique genes. Interestingly, using the same placental samples for the transcriptomic analysis, the authors identified 163 differentially expressed genes (DEGs) among the targeted genes by DMCs: 124 were downregulated and 39 were upregulated between women with a high-level exposure of phthalates compared to those with a low-level exposure. The authors focused on one particular pathway, erbB signaling. EGFR exhibited placental hypermethylation and decreased expression in women, with a high total phthalate exposure, suggesting that this gene might be the specific target of endocrine disruption by phthalates. Alterations in the erbB signaling methylome have previously been observed in adverse obstetric outcomes such as pre-eclampsia and IUGR.
In two recently published studies [90,91] based on a larger cohort of mother–infant pairs (the French EDEN cohort, N = 202), placental DNA methylation at birth was measured using the Illumina Infinium Human Methylation 450K Bead Chip in combination with 12 urinary phthalate metabolites [91] and 9 phenols [90]. In these studies, only placental samples from boys were assessed for DNA methylation. Since biological functions are in general more strongly associated with genomic regions than single CpGs [92], these two studies aimed to identify DMRs associated with phenol and phthalate concentrations. Most of the associations observed were positive (i.e., increased methylation with increased exposure during pregnancy). TCS exposure was positively associated with 37 DMRs compared to less than 6 for the other phenols and phthalate metabolites studied. Out of the 37 DMRs associated with TCS, 6 encompassed imprinted genes, which represented a significant enrichment. Using the SEPAGES cohort (N = 387 mother–child couples), an exploratory analysis on individual CpGs and DMRs, as well as 20 previously identified CpGs, was performed. In the sex-stratified analysis, 114 individual CpGs (68 in males and 46 in females) were differentially methylated, encompassing 74 genes (36 for males and 38 for females). There was a total of 82 significant DMRs (40 for females and 42 for males). For most DMRs, DNA methylation levels increased with a higher exposure, except for some parabens and DiNP metabolites in males and females, BPA in males, and BPS and DiNCH metabolites in females, which were negatively associated with DNA methylation in most DMRs. Some differentially methylated CpGs and DMRs encompassed imprinted genes, whereas other CpGs were linked to adiposity, the lipid and glucose metabolism, and cardiovascular function [93]. Recently, maternal concentrations of monocarboxyisononyl phthalate (cx-MiNP), mono-3-carboxypropyl phthalate (MCPP), and BPA were associated with altered methylation in the placenta (the maternal or fetal side). Among them, MCPP was associated with differential CpG methylation [94].
Further studies are needed to better control the analysis conditions (the PPP exposure time, placental sampling, pregnancy stages, genotyping, etc.) and to acquire more causal information between the methylation landscape, gene expression, and placental dysfunction induced by EDCs.
5.3. Effects of PPPs on Human Placental Function Using In Vitro Models
In humans, trophoblasts differentiate into extravillous cytotrophoblasts that proliferate, invade the uterine wall, and are important for remodeling the endometrial vasculature and syncytiotrophoblasts. The latter cell type resulting from the fusion of cytotrophoblasts is responsible for the endocrine function of the placenta that produces steroid (progesterone) and polypeptide hormones (human chorionic gonadotropin, namely, hCG, and leptin). To study the effects of PPPs on the human placentation, in vitro studies could be performed using (i) primary cell cultures of trophoblasts from first- or third-trimester placentae [95,96,97,98,99], (ii) placental explants [100,101], and (iii) immortalized trophoblast cell lines such as BeWo, JEG-3, and HTR-8/SVneo [102,103,104,105]. The BeWo cell line can fuse and express differentiation markers [106,107,108,109], whereas JEG-3 cells can produce placenta-specific hormones, but without cell fusion [110,111]. The HTR-8/SVneo cell line is used as a model of extravillous trophoblasts [112,113,114].
Using human placenta, the exposure of first-trimester placental explants to BPA (0.5–1 nM) stimulated hCG secretion [115,116] and increased cell apoptosis [116]. The same effects were observed when term primary trophoblasts were exposed to BPA (8 nM–8 mM) [117,118]. Unlike the first-trimester explants, Zou et al. did not observe any change in the level of hCG secretion at any BPA concentration tested in term villous explants. Interestingly, they showed that BPA significantly increased ESRRG expression (estrogen-related receptor gamma) in the female placentae following exposure to 1 µM BPA for 24 h. This suggests that exposure to a low dose of BPA could alter gene expression in human placentae in a sex-specific manner [100]. Moreover, in primary trophoblast cells, data have shown that BPA is also able to (i) induce the expression of corticotrophin-releasing hormone (CRH) and the expression of two enzymes specifically involved in hormone production (aromatase and 11-β-hydroxysteroid dehydrogenase 2, known as 11β-HSD2), and, conversely, (ii) to reduce the expression of leptin obtained in human term primary trophoblasts [117]. Concerning BPS, this phenol has been shown to block epidermal growth factor (EGF)-mediated trophoblast fusion in term primary trophoblasts [118].
In the first-trimester trophoblast progenitor cells and second-trimester primary trophoblasts, a mixture of four phthalates (MBP, 0.2 µM; MBzP, 3 µM; MEHP, 0.7 µM; MEP, 1.5 µM) decreased hCG expression regardless of the fetal sex. However, the same mixture of phthalates decreased the placental peroxisome proliferator-activated receptors-γ (PPARγ) expression in male cells and increased female cells. This latter finding provides evidence for sex-specific responses to phthalates in human trophoblasts [98]. In primary-term trophoblasts, a recent study demonstrated that, at low concentrations (0.1–1 µM), MEHP decreases the lipid content, hCG secretion, and cell fusion. These effects appear to be associated with a lower activity of (PPARγ). This receptor is a key transcription factor involved in the control of trophoblast differentiation and the lipid metabolism. In contrast, a high concentration (10 µM) of MEHP increases the lipid content, cell fusion, and PPARγ activity, but decreases the hCG secretion. These results highlighted the notion of a non-monotonic dose–response, particularly at environmentally relevant levels of MEHP exposure [97]. Moreover, at this stage, high MEHP doses (100–150 µM) increase the expressions of CRH and cyclooxygenase 2 (COX2) via the nuclear factor κ-light-chain-enhancer of activated B cells (NF-κB) signaling pathway. These results suggest that MEHP could prematurely induce the expression of labor-promoting genes and lead to preterm birth [119].
Using the BeWo cell line, exposure to low doses of BPA (3 nM and 1 µM) has been shown to significantly increase cell proliferation [120,121], whereas exposure to a high dose of BPA (1000 µM) decreased the proliferation rate of cells [121,122]. However, in other studies, the exposure of BeWo cells to BPA did not affect the proliferation or metabolic activity [123]. BPA exposure has also been shown to induce cell cycle inhibition, increase DNA damage [122], and reduce the viability of BeWo cells at concentrations ranging from 100 µM to 1 mM [116]. Regarding β-hCG secretion, a significant decrease was observed at a 30 µM BPA exposure, while the secretion dropped dramatically from 60 to 125 µM BPA [116]. BPA is able to promote the syncytialization in BeWo cells at doses of 1 and 50 mM, known in humans to increase the expression of several syncytin proteins [123], as well as to reduce the invasive capacity of the same cell line at doses as low as 100 µM [124] due to the increased expression of E-cadherin. BPA was also able to reduce the activation of the antioxidant response element [121] and increase the levels of two anti-apoptotic proteins, i.e., B cell lymphoma 2 (Bcl-2) and heat shock protein 70 (Hsp70), and decrease the levels of hypoxia-inducible factor 1-α (HIF-1α) under stress conditions, demonstrating that BPA inhibits trophoblast cell death under cellular stress conditions [121]. Moreover, BPA might exert its various effects mainly by regulating cell–extracellular matrix interactions via transforming growth factor-β (TGF-β) signaling [125]. Cytotoxic effects via cell cycle arrest and apoptotic pathways were also observed in BeWo cells exposed to EtP [126].
As JEG-3 cannot fuse, there are few studies on this cell line. It has only been shown that, at doses of 5–20 µg/mL, BPF induces chromatin condensation [102,103,104]. Concerning MEHP exposure, the incubation of JEG-3 cells (10 mM for 72 h) exhibited a marked change in the lipid profile, especially in triacylglycerols and glycerophospholipids, with a marked accumulation of triacylglycerols. This once again underlines the detrimental effects of MEHP on the lipidome of human placental cells [127].
Using the HTR-8/SVneo cell line, a high dose of TCS exposure suppressed the viability and migration of HTR-8/SVneo cells but increased the H19 imprinted gene expression and insulin-like growth factor 2 (IGF2) protein secretion in a dose-specific manner. This suggests an altered trophoblast function, gene expression, and DNA methylation status [112]. Regarding the exposure to the paraben family, n-BuP has also been shown to inhibit cell proliferation and induce apoptosis, which could disrupt early placental development [128]. Regarding phthalate exposure, high concentrations of MEHP for 24 h significantly inhibited cell proliferation and viability, promoted apoptosis, and inhibited the cell cycle [129]. In addition, DEHP acts on placental trophoblast cells and inhibits its internalization of transthyretin (thyroxine transport protein), downregulates transthyretin expression, and affects the expression of deiodinase 2, deiodinase 3, and thyroid hormone receptor in the HTR-8/SVneo cell line, as well as in JEG-3 cells [130]. These data suggest that DEHP disrupts the placental transport of thyroid hormones, which could be very detrimental to fetal development, especially at early stages.
In conclusion, depending on the phenol concentrations, BPA may or may not affect placental cell proliferation and hormone production. Parabens, on the other hand, induce cytotoxic effects and apoptosis. Depending on their concentration, phthalates affect the lipid content, hCG secretion, and cell fusion by PPARγ. All of these cell models are useful for testing several concentrations of PPPs and deciphering their mechanism of action. However, these approaches do not allow us to take into account the sub-chronic effects, nor the complexity of the different cell types of the placenta. It is, therefore, essential to supplement the in vitro data with data from in vivo models.
5.4. Effects of Phenols and Parabens on Fetoplacental and Neonatal Outcomes Using Animal Models (Table 2)
As a very detailed review of phthalate exposures and placental health has been published by Seymore et al. in 2022 [131], an update has been made over the last two years. Exposure to 20 µg/kg BW/day of DiNP on days 1–7 of gestation in mice led to fetal loss towards the end of gestation. This miscarriage was due to the impaired differentiation of stromal cells associated with a defect in the angiogenic network in the decidua, but also to a disorganization of the placental layers. Interestingly, the labyrinthine area involved in nutrient exchanges was reduced [132].
In rats, oral exposure to BPA (0.0025, 0.025, and 0.250 mg/kg/day) 30 days before pregnancy and during the first 20 days of pregnancy reduced the placental weight at low and high BPA concentrations, while the fetal weight was increased only at low concentrations [133]. In contrast, gestational exposure to BPA in ewes reduced the placental efficiency and fetal weight at mid-gestation [134] and at 110 days [135]. Similarly, the exposure of pregnant mice to BPA at 0.050 mg/kg/day by oral gavage from day 1 to 7 of gestation induced fetal growth restriction at 14 days of gestation, which was preceded by the insufficient remodeling of uterine spiral arteries [136]. In addition, a significant decrease in the number of embryos and the weight of the uterus on days 10 and 12 was reported after the subcutaneous administration of BPA (10 mg/kg/day) to pregnant mice from day 1 of gestation to day 7 [137].
The study of the placental structure was analyzed after exposure to BPA. Tait et al. exposed pregnant mice to BPA at 0.5 mg/kg bw (corresponding to the lowest observed adverse effect level, LOAEL) or 50 mg/kg bw (corresponding to a high range of human BPA exposure) from GD1 to GD11. They demonstrated that a high concentration of BPA induces a significant degeneration and necrosis of giant cells, associated with an increase in vacuolation in the junctional zone and a reduction in the spongiotrophoblast layer at GD12 [138]. Degenerative changes were also found in the trophoblastic giant cells and spongiotrophoblast layer in the pregnant mice exposed to BPA (10 mg/kg/day, from day 1 to day 7) by subcutaneous administration [137]. A low concentration of BPA (0.5 mg/kg bw) induced the development and branching of blood vessels, while a high concentration of BPA inhibited them. Maternal vessels were narrower in placentae exposed to a low concentration of BPA, whereas embryonic and maternal vessels were irregularly dilated in the labyrinthine area of placentae exposed to a high concentration [138]. Doses of 10 and 40 mg/kg BPA (mice exposed from GD0.5 to GD5.5) also decreased the proportion of the labyrinthine and spongiotrophoblast layers, which are associated with large vacuoles [139]. An increase in the retention of smooth muscle cells and a decrease in vascular surfaces at the level of the junctional zone have been reported in mice exposed to BPA at a concentration of 0 to 400 µM in drinking water from GD7 to GDl7 [140]. Altogether, these data indicate that BPA has an effect on the placental structure, which explains the disturbance to the fetal weight.
BPA-related gene expression changes have also been studied. Since BPA is thought to bind to estrogen receptors and regulate the expression of estrogen-responsive genes, BPA may alter the expression of other placental nuclear receptors such as the retinoid Z receptor (RORγ), progesterone receptor (PR), ERβ, LXRα, germ cell nuclear factor (GCNF), steroidogenic factor-1 (SF-1), and so on. The oral administration of BPA at a dose of 0.002 mg/kg/day from GD6.5 to GD17.5 induced the sex-specific placental differences in these genes at GD18.5 [141].
Post-translational modifications of histones, DNA methylation, and the expression of imprinted genes have been studied in mouse placenta concerning BPA exposure. Daily exposure of pregnant mice to BPA between GD8.5 and GD12.5 affected the expression of imprinted genes. At GD13.5, Rtl1 displayed a slight disturbance of allele-specific expression [142]. A longer maternal exposure time to BPA (two weeks before mating and during gestation) significantly disrupted the expression of imprinted genes in the placenta at GD12.5, such as Snrpn and Kcnq1, and altered the methylation levels of differentially methylated regions (DMRs), including the Snrpn imprinting control region (ICR) and Igf2 DMR1 [143] Moreover, exposure significantly reduced the genome-wide methylation levels in the placenta, but not in the embryo. The expression of several small RNAs was also found to be disturbed [144]. These epigenetic defects were associated with abnormal placental development [143]. The exposure of pregnant mice from GD7 to GDl7 to BPA at a concentration of 0 to 400 µM in drinking water disrupted the expression of 10 genes coding the proteins of epigenetic machinery. These genes include enzymes involved in histone methylation and acetylation, protein phosphorylation, and DNA methylation. Among these enzymes, the DNMT1 mRNA/protein was increased in the placenta, while levels of DNMT3a and -3b, hydroxymethyl transferases (TET1, -2, and -3), and 5-hmc were unaffected. Moreover, high levels of 5-mc indicated an elevated level of methylation in the placental tissues of mice exposed to BPA [140].
As previously described, BPA altered the placental structure, leading to impaired placentation. This could result from increased protein levels of matrix metalloproteinase-9 and 2 (MMP-9 and MMP-2) and decreased levels of the tissue inhibitor of metalloproteases-3 (TIMP-3), as well as integrin-β1 and integrin-α5, in mouse placenta. These effects were observed in particular in the labyrinthine layer in the event of a low exposure to BPA (5, 10, and 40 mg/kg) from GD0.5 to GD5.5 [139]. Ye et al. confirmed that placental abnormalities were associated with altered invasion-related genes, such as increased tissue inhibitors of metalloproteinases, decreased metalloproteinases, and the Wnt family member WJVT2/β-catenin (mice exposed from GD7 to GDl7 to BPA added at a concentration of 0–400 µM in drinking water) [140]. BPA affected several placental transporters, such as cation transport channels. For example, in mice placenta at GD17.5, PMCA1 (ATPase, Ca++ transporting, plasma membrane 1), hephestin (HEPH), CTR1 (solute carrier family 31, member 1 (copper transporter)), and ATP7A (ATPase, Cu2+ transporting, alpha polypeptide) were disturbed after exposure to BPA (50 mg/kg/day) from GD11.5 to GD16.5. These disturbances were associated with a decrease in serum calcium/copper/iron levels, which could have an impact on fetal development [145]. Exposure to high doses of BPA (400 and 600 mg/kg) from GD17 to GD19 increased CaBP-9k mRNA/protein in the maternal uterus and placenta in late gestation. CaBP-9k is a vitamin D-dependent calcium-binding protein whose gene carries an estrogen response element (ERE) [146,147]. In rats fed a diet containing BPA (0.0025, 0.025, or 0.250 mg/Kg/day) for a month, plus 20 days during pregnancy, the glucose type 1 transporter was upregulated [148]. BPA modulation of cation transport channels, CaBP, and placental nutrient-glucose transfer could explain the changes in the fetal weight.
As exposure to BPA in pregnant mice (2, 20, and 200 mg/kg bodyweight/day from GD13 to GD16, and euthanized at GD17) increased the plasma estrogen, testosterone, and CRH in dams, the regulation of CRH involved in fetal organ development, the glucose metabolism, and immune response has been investigated in the placenta. The activation of phosphorylated forms of PKC ζ/λ and δ might promote cAMP-responsive element-binding protein (CREB) phosphorylation. This leads to its interaction with a CBP-responsive element (CRE) located in the CRH gene promoter and to an increase in CRH mRNA, which might be the pathway for the signaling of preterm birth [149].
Other mediators differ between mid- and early gestation in terms of protein expression in the BPA-exposed placenta. Exposure to BPA in ewes at the start of gestation increased interleukin 8, the marker of lipid peroxidation, antioxidants, aromatase, 17 alpha-hydroxylase, estrogen receptor 2, IGF-2 receptor and IGF-binding proteins (IGFBPs), and histone deacetylase 1 and 2, and caused a reduction in tumor necrosis factor-alpha and the IGF1 receptor. Whereas, at mid-gestation, BPA exposure reduced angiogenic factor hypoxia-inducible factor 1 alpha, but increased IL1beta, oxidative stress markers, triglyceride, 17-alpha hydroxylase, IGFBP 1, DNA methyltransferase 3 A, and histone deacetylase 1, which could explain the low birth weight [134]. Exposure to BPA from day 40 through day 110 of gestation in ewes (5 mg/kg/d) was responsible for the placental cytotoxicity, including autophagy, apoptosis, endoplasmic reticulum stress, excessive ROS generation, oxidative damage, and mitochondrial dysfunction [135].
BPA is the most studied phenol so far, but other molecules of the same family are also studied, including BPS, TBBPA, and TCS. BPS is an analog of BPA that is reputedly more inert. Daily exposure to BPS from GD30 to 100 in pregnant ewes reduced maternal circulating pregnancy-associated glycoproteins (PAG1 and PSPB) but did not change the placental weight or placental stereology [150]. However, exposure to BPS in mice (0.200 mg/kg bodyweight BPS or BPA 2 weeks before mating and until day 12.5 of gestation) was shown to reduce the ratio of the spongiotrophoblast zone to trophoblast giant cells within the junctional zone, as well as the exposure to BPA. In addition, BPA and BPS altered placental neurotransmitters such as serotonin and dopamine. BPA and BPS reduced placental serotonin (5-HT) concentrations and 5-HT giant cell immunoreactivity, whereas the concentrations of dopamine and 5-hydroxy indole acetic acid, the main metabolite of serotonin, increased, as well as the dopamine immunoreactivity of the giant cells. Due to BPA and BPS exposure, this imbalance associated with a decrease in docosahexaenoic acid and estradiol could affect the placental–brain axis of the mouse fetus. BPS exposure causes placental effects almost identical to those of BPA, which would justify considering BPS as being as dangerous as BPA [151]. Finally, placentae exposed to BPS in pregnant ewes (from day 30–100) showed a low expression of the protein E-cadherin, few binucleate cells, and a high expression of missing-1 protein in glial cells, suggesting that BPS can affect the trophoblast fusogenic signaling pathway and the placental endocrine function [150].
The administration of TCS at doses of 523 and 785 mg/kg/day on GD1 to GD3 has been observed to impair blastocyst implantation in mice [152]. In animal models, TCS was able to bioaccumulate in the placenta, liver, kidney, ovary, adrenal, spleen, and fat, but with high concentrations for the first four tissues [153,154]. Daily oral exposure of pregnant mice with 8 mg/kg of TCS, but not 1 or 4 mg/kg, from GD6 to GD18 resulted in a decreased fetal bodyweight and an increased rate of fetal loss (spontaneous abortions and fetal growth restriction) [155]. Similarly, a short exposure to TCS from GD5.5 to mid-gestation caused a dose-dependent increase in the rate of fetal loss through abortion [156]. Spontaneous abortion has also been reported in pregnant rats exposed to TCS by gavage with 600 mg/kg/d from GD6 to GD20 [154]. A high TCS concentration decreased the placental weight (exposure by gavage at doses of 1, 10, and 100 mg/kg/day from GD5.5 to GD15.5), and the placental structure was characterized by a thrombus, hemorrhage with tissue necrosis, and junctional zone atrophy [156]. Exposure to 8 mg/kg/day from GD6 to GD18 confirmed the reduction in placental weight and labyrinth volume [155]. Taken together, these data demonstrated that TCS is capable of inducing miscarriage and affecting the placental structure. In addition, TCS negatively modulated the activities and expression of placental System A amino acids or glucose transporters in pregnant mice exposed to 8 mg/kg/day. These negatively modulated activities were associated with a decrease in the sodium-coupled neutral amino acid transporters (SNAT1/SNAT4) and glucose transporter 1 (GLUT-1) mRNA, respectively [155]. These indicate an effect on nutrient exchanges that may contribute to fetal growth restriction.
Since TCS is an endocrine disruptor, the levels of reproductive hormones and thyroid hormones, and their enzymes, have also been studied. In pregnant ewes, as well as in mice, a high concentration of TCS decreased estrogen sulfonation [153]. Indeed, estrogen sulfotransferase activities, implicated in both estradiol and estrone sulfonation, were significantly reduced when pregnant mice were exposed to TCS. Although the serum estrogen concentration was normal, the ratio of sulfo-conjugated E2 and unconjugated E2 was reduced in mice exposed to TCS. Interestingly, the estrogen receptor antagonist, an estrogen sulfotransferase activity inhibitor, was able to rescue the platelet aggregation and placental thrombosis, and limit spontaneous abortion [156].
Feng et al. studied the impact of TCS exposure on placental steroid metabolism enzymes, including UDP-glucuronosyltransferase 1A1 (UGT1A1), estrogen sulfotransferase 1E1 (SULT1E1), and steroid 5α-reductase 2 (SRD5A2). This study indicated a significant impairment with high concentrations of TCS as well as progesterone and estrogen receptors. This could explain why placental hormones such as progesterone, estradiol, testosterone, and prolactin secreted in maternal blood were reduced in groups exposed to high doses of TCS [154]. T3 and T4 are known to affect fetal growth and development. In the event of exposure to a high concentration of TCS (8 mg/kg/day orally from GD6 to GD18), TCS induced hypothyroxinemia in pregnant mice. As thyroid hormones stimulate the Akt-mTOR-p70S6K and ERK signaling pathways, which can regulate the activation of placental amino acid transporters, these pathways have also been investigated. TCS decreased placental Akt, mTOR, and P70S6K phosphorylation, but this was corrected by L-thyroxinein (T4). In fact, T4 was able to rescue the activity and expression of amino acid and glucose transporters and decrease fetal bodyweight [155].
In pregnant rats, following the administration to dams of 100, 200, and 400 mg/kg bodyweight/day from GD7 to 21, parabens (EtP and n-BuP) were distributed in rat maternal plasma, pools of amniotic fluids, placenta, whole-body fetuses, and in the fetal liver. Additionally, high levels of EtP were found in all fluids and tissues compared to n-BuP [157].
All of these experiments relating to maternal oral exposure to PPPs and placental effects were always performed using one pollutant, variable doses, and different timings of exposure, which rarely covered the preconceptional and gestational periods.
Table 2Effects of phenols, phthalates, and parabens on fetoplacental development according to the animal model, the chemicals, the dose administrated, the exposure route and duration.
| Animal Model | Chemicals | Dose Administered | Exposure Route | Exposure Duration | Observation Stage | Function Studied | Fetoplacental Outcomes | Additional Outcomes | Reference |
|---|---|---|---|---|---|---|---|---|---|
| Mouse | BPA | 0.002 mg/kg bw/d | Oral gavage | GD6.5-17.5 | GD18.5 | Placental function | Peroxisome proliferator-activated receptor alpha and gamma (PPARα, PPARγ) and Aryl hydrocarbon receptor (AhR) mRNA expression: no effect | COUP-TFα, GCNF, SF-1, and PNR mRNA expression: BPA ↓ (♂)/↑ (♀) | [141] |
| Mouse | BPA | 0.05 mg/kg bw/d | Oral gavage | GD1-7 | GD5, 8, 10, 12, 14 | Fetoplacental development/placenta and uterine spiral artery structure | Placental areas, thickness, and diameter: no effect | Implantation number and abortion rates: no effect | [136] |
| Mouse | BPA | 0.5 or 50 mg/kg bw/d | Oral gavage | GD1-11 | GD12 | Fetoplacental development/placental structure | Placental diameter and embryo length: BPA ↑ (50) | [138] | |
| Mouse | BPA | 0, 2, 20, or 200 mg/kg bw/d | Oral gavage | GD13-16 | GD17 | Placental function | Placental CRH mRNA expression: BPA ↑ (200) | Plasma CRH, estradiol, and testosterone concentrations: BPA ↑ (20, 200) | [149] |
| Mouse | BPA | 0, 0.4, 4, 40, or 400 µM BPA in drinking water | Orally (drinking water) | GD7-17 | GD13, 16 and 17 | Placental structure and function | Mean fetal and placental weight, litter size, and placental-to-bodyweight ratio: no effect | Dams systolic blood pressure: BPA ↑ (4, 40, 400) | [140] |
| Mouse | BPA | 200 mg/kg bw/d | Orally (food) | GD-15 to GD12.5 | GD12.5 | Placental function | Placental number of differentially expressed miRNAs: BPA ↑ (22)/↓ (21) | [144] | |
| Mouse | BPA | 10 mg/kg bw/d BPA | Subcutaneous injections | GD0-GD7 | GD10, GD12, PND0-56 | Placental structure | Number of embryos and survival rate: BPA ↓ | Uterine weight, uterine-to-bodyweight ratio, and glycogen-containing cells: BPA ↓ | [137] |
| Mouse | BPA | 5, 10, or 40 mg/kg bw/d | Subcutaneous injections | GD0.5-5.5 | GD14.5 | Placenta structure and function | Labyrinthine area and decidua ITGβ1 and ITGα5 protein levels: BPA ↑ | - | [139] |
| Rat | BPA | 0.0025, 0.025, or 0.25 mg/kg bw/d | Orally (drinking water) | GD-30 to GD20-22 | GD20-22 | Fetoplacental development and uterine artery | Fetal weight: BPA ↑ (2.5) | Uterine artery diameter and uterine artery relaxation: BPA ↓ | [133] |
| Rat | BPA | 0.0025, 0.025, or 0.25 mg/kg bw/d | Orally (drinking water) | GD-30 to GD20 | GD20 | Fetoplacental development/placental metabolism | Fetal weight at birth: BPA ↑ (2.5) | Non-pregnant rats bodyweight: BPA ↑ (25, 250) | [148] |
| Rat | BPA | 200, 400, or 600 mg/kg bw/d | Subcutaneous injections | GD17-19 | PND5 | Maternal and fetal uteri gene expression | Postnatal uterine calbindin-D9K (CaBP-9k) and Estrogen receptor alpha (Erα) mRNA expression: ↑ (600) | Maternal uterine CaBP-9k mRNA expression: BPA ↑ (600) | [147] |
| Sheep | BPA | 0.5 mg/kg bw/d | Subcutaneous injections | GD30-90 | GD65, 90 | Placental function | Fetal weight and placental efficiency: BPA ↓ (GD65) | [134] | |
| Sheep | BPA | 5 mg/kg bw/d | Subcutaneous injections | GD40-110 | GD110 | Placental function/fetal development | Placentome total weight, fetal weight, and placental efficiency: BPA ↓ | [135] | |
| Mouse | BPA, BPS | 200 mg/kg bw/d | Orally (food) | GD-15 to GD12.5 | GD12.5 | Placental structure, function, and metabolism | Number of fetuses and placental oestrone, corticosterone, testosterone and progesterone (P4) concentration: no effect | Maternal gestational weight, success of pregnancy, and number of implantation sites: no effect | [151] |
| Sheep | BPA, BPS | 0.5 mg/kg bw/d | Subcutaneous injections | GD30 to GD100 | GD120 | Placental function | Endocrine function (pregnancy-associated glycoprotein 1 (PAG1), pregnancy-specific protein B (PSPB), P4): BPS ↓ | [150] | |
| Rat | BPA, DEHP | 0.005 mg/kg bw/d BPA, 5 or 7.5 mg/kg bw/d DEHP, or both (mix) | Oral gavage | GD6-21 | GD6 (abortion and weight), PND1 to 24 weeks | Fetal development | Chest circumference: DEHP ↑ (5)/DEHP + BPA ↑ (5)/DEHP ↓ (7.5)/DEHP + BPA ↓ (7.5) | Abortion rate: BPA + DEHP ↑ | [158] |
| Mouse | BPA, OP | 50 mg/kg bw/d | Subcutaneous injections | GD11.5-16.5 | GD17.5 | Placental function | Placental calcium transporter channels: OP ↓ (Trpv6 protein level)/OP ↑ (Pmca1 mRNA and protein levels), BPA ↑ (Pmca1 mRNA and protein levels) | [145] | |
| Rat | BPA, OP, NP | 200, 400, or 600 mg/kg bw/d | Subcutaneous injections | GD20 | Placental function | Placental CaBP-9k mRNA expression: BPA ↑ (200) | Maternal uterus CaBP-9k mRNA expression: OP/NP ↑ (600) | [146] | |
| Mouse | BPA, TBBPA | 10 mg/kg bw/d BPA, 0.5 mg/kg bw/d TBBPA | Orally (food) | GD-15-GD16.5 | E6.5-E10.5, E16.5 | Fetoplacental development/placental function | Rate of the hemorrhaging conceptus and fetal loss: BPA ↑/TBBPA ↑ | Resorption rate: BPA ↑/TBBPA ↑ | [159] |
| Mouse | BPA, TCS | 0, 87, 262, 523, or 7858 mg/kg bw/d TCS in a single dose; 0, 523, or 7858 mg/kg bw/d TCS in repeated doses; 61 or 122 mg/kg BW/d BPA; 262 mg/kg BW/d TCS + 61 mg/kg BW/d BPA; 262 mg/kg BW/d TCS + 122 mg/kg BW/d BPA | Subcutaneous injections | GD1-3 | GD6, 17, PND 0, 4, 7, 14, and 21 | Implantation | Number of implantation sites: TCS ↓ (repeated-dose GD6: 523, 7858)/TCS ↓ (single-dose GD2: 7858/GD3: 523, 7858)/TCS + BPA ↓ (GD6: 262 + 122) | [152] | |
| Mouse | TCS | 1, 4, or 8 mg/kg bw/d | Oral gavage | GD6-18 | GD11-19 | Fetoplacental development/placental function | Number of live fetuses, fetal bodyweight, and placental weight: TCS ↓ (8) | Triiodothyronine (T3), thyroxine (T4) and P4 levels: TCS ↓ (8) | [155] |
| Mouse | TCS | 4 or 8 mg/kg bw/d | Oral gavage | GD6-14 | PND1, 30, 60 | Placenta/metabolism | Numbers of pups, offspring bodyweight, and body mass index: TCS ↓ (8) | [160] | |
| Mouse | TCS | 0, 10, 50, or 100 mg/kg bw/d | Oral gavage | GD7.5-17.5 | GD17.5 | Placental function | Fetal and placental weight: TCS ↓ (50, 100) | Uterus size: TCS ↓ (50, 100) | [161] |
| Rat | TCS | 75, 150, or 300 mg/kg bw/d | Oral gavage | GD8-PND21 | PND3, PND90 | Fetal development/uterine structure/estrous cycle | Mean of litter weight: TCS ↓ (PND3) | Maternal bodyweight: TCS ↓ | [162] |
| Sheep | TCS | 0.1 mg/kg bw/d | Direct infusion into the fetal circulation or through administration to the ewe | 2 days between GD120 and GD130 | 3 days after 1st injection | Placental/liver sulfotransferase activity | TCS concentration: Placental concentration > Fetal liver concentration | [153] | |
| Mouse | DEHP | 0, 50, or 200 mg/kg bw/d | Oral gavage | GD0-6, GD7-12, GD13-17 | GD18 | Fetoplacental development | Fetal weight: DEHP ↓ | [163] | |
| Mouse | DEHP | 125, 250, or 500 mg/kg bw/d | Oral gavage | GD1-13 | GD9 and GD13 | Placental development and function | Embryo implantation: DEHP ↓ (500) | Protein levels involved in apoptosis process: DEHP ↑ (GD13: Bax casp-3 and Bax caps-8)/↓ (GD13: Bcl2) | [164] |
| Rat | DEHP | 500 or 1000 mg/kg bw/d | Oral gavage | GD7-12 | GD20 | Fetal development/placental function | Number of living fetuses: DEHP ↓ | Pregnant weight changes at GD18: DEHP ↓ | [165] |
| Rat | DEHP | 750 or 1500 mg/kg bw/d | Oral gavage | GD0-19 | GD20 | Placental metabolism | PPARα and PPARγ mRNA expression in the labyrinth and the junctional zone: DEHP ↑ | Arachidonic acid (AA) in fetal and maternal plasma: DEHP ↓ | [166] |
| Rat | DnHP, DCHP | 0, 20, 100, or 500 mg/kg bw/d | Oral gavage | GD6-19 | GD20 | Fetoplacental development | Fetal bodyweight: DnHP ↓ (20, 100)/DnHP ↑ (♂: 500)/DCHP ↓ (♀: 500)/DCHP ↑ (20, 100) | F0 relative kidney and liver weights: DCHP ↑ (20)/DnHP↓ (500) | [167] |
| Rat | n-BuP | 0, 10, 100, or 1000 mg/kg bw/d | Oral gavage | GD6-19 | GD20 | Fetal development | Litter size, survival rate, number of resorptions, sex ratio, and fetal bodyweight: no effect | Dams bodyweight and bodyweight changes: no effect | [168] |
| Mouse | n-BuP, PrP | 0, 0.05, 0.5, 5,10, 20, 30, and 35 mg/dams n-BuP; 35 or 40 mg/dams PrP | Subcutaneous injections | GD1-4 | PND0, 3, 5 | Fetal development | Skin irritation after injection: n-BuP ↑ (0.5, 5, 10, 20, 30, 35) | [169] | |
| Rat | DiBP, BP | 600 mg/kg bw/d DiBP or 100 mg/kg bw/d n-BuP | Oral gavage | GD7-19 or GD7-21 | GD19-21 | Fetal metabolism | Plasma leptin level: DiBP ↓/n-BuP ↓ | Dams bodyweight gain: no effect | [170] |
| Rat | EtP, n-BuP | 100, 200, or 400 mg/kg bw/d | Subcutaneous injections | GD7-21 | GD21 | Placental function | Number of fetuses, fetus weight, and fetus organ weight: no effect | Maternal weight: no effect | [157] |
| Mouse | EtP, PrP | 0, 400, 800, or 1600 mg/kg bw/d EtP, and 0, 625, 1250, or 2500 mg/kg bw/d PrP | Oral gavage | GD1 to GD5-7 or to GD8-9 | GD5, | Implantation/endocrine function | Marker expression of decidualization: EtP ↓ (1600)/PrP ↓ (2500) | Implantation rate: EtP ↓/PrP↓ | [171] |
Results in brackets correspond to the sex in which the effect was observable (♂, male; ♀, female), followed by stage at observation separated by a column from the doses administered at which the effect was observed, and, if so, followed by the number of DEGs in brackets. If one of the above-mentioned characteristics is not specified, the effect was observed for both sexes at the different stages of observation and the different doses used. Arrows indicate the direction of the effect: a downward arrow (↓) indicates a decrease and an upward arrow (↑) indicates an increase. ANGPTL4: angiopoietin-like protein 4; Ascl2: achaete-scute family bHLH transcription factor 2; ATP: adenosine triphosphate; ATP7A: ATPase Cu++ transporting alpha-polypeptide; BPA: bisphenol A; BPS: bisphenol S; n-BuP: butylparaben; bw: bodyweight; cAMP: cyclic adenosine monophosphate; CD3: cyclin D3; Crt1: solute carrier family 31: member 1; COUP-TFα: chicken ovalbumin upstream promoter transcription factor alpha; COX-2: cyclooxygenase-2; CRH: corticotropin-releasing hormone; CYP19: placental aromatase; CYP17: 7 alpha hydroxylase; CYP4A1: cytochrome P450 subfamily 4A1; d: day; DCHP: dicyclohexyl phthalate; DEHP: di(2-ethylhexyl) phthalate; DnHP: di-n-hexyl phthalate; DiBP: diisobutyl phthalate; DNMT3A: DNA methyltrasnferases; Eomes: eomesodermin; EtP: ethylparaben; ERK: extracellular signal-regulated kinase; ESR2: estrogen receptor 2; Esx1: extraembryonic, spermatogenesis, homeobox 1; F1: first generation; FAT/CD36: fatty acid translocase; FATP1: fatty acid transport protein 1; Fosl1: Fos-like antigen 1; GCM1: glial cell missing factor 1; GCNF: germ cell nuclear factor; GD: gestational day; Hand1: heart and neural crest derivatives-expressed 1; HDAC: histone deacetylases; Heph: hephestin; HFABP: heart cytoplasmic fatty acid-binding protein; Ido1: indoleamine2,3 deoxygenase1; IGF1R: insulin-like growth factor 1 receptor; IGF2R: insulin-like growth factor 2 receptor; IGFBP: insulin-like growth factor-binding protein; IL-1β: interleukin 1 beta; IL-8: interleukin 8; LXRα: liver X receptor alpha; MAPK: mitogen-activated protein kinase; MeAIB: methylaminoisobutyric acid; miRNAs: microribonucleic acids; MMP: metalloproteinase; mRNA: messenger ribonucleic acid; NP: 4-nonylphenol; OP: 4-tert octylphenol; PCNA: proliferating cell nuclear antigen; PMCA1: plasma membrane Ca2+ ATPase; PND: postnatal day; PNR: photoreceptor-specific nuclear receptor; PrP: propylparaben; SF-1: steroidogenic factor-1; SNAT: sodium-coupled neutral amino acid transporter; TBBPA: tetrabromobisphenolA; TCS: triclosan; TIMP: tissue inhibitor of metalloproteinase; TNF: tumor necrosis factor alpha; Tregs: regulatory T cells; Trpv6: transient receptor potential cation channels in subfamily V member 6; VEGF: vascular endothelial growth factor; WNT: Wnt family member.
6. Effects of PPPs on Offspring Health after Birth from Epidemiological Data and Using Animal Models
Guilbert et al. and Radke et al. suggested that PPPs impact child neurodevelopment (including cognition and behavior) [39,54]. In this review, only the effects of PPPs in terms of obesity, metabolic alterations, cardiovascular disorders, and fertility defects will be explored.
6.1. Effects of PPPs on Obesity
6.1.1. Epidemiological Data (Table 3)
Few studies have examined the effects of prenatal PPP exposure on postnatal adiposity and growth [113,172,173,174,175]. Regarding phenols, we identified 16 studies from nine countries (Canada, China, Denmark, France, Greece, Mexico, Republic of Korea, Spain, and the USA) that examined the association between maternal phenol concentrations measured in urine and postnatal weight and adiposity in children. Sample sizes ranged from 218 to 1301 [173,175,176]. The outcomes measured varied across studies, with weight, BMI, waist circumference, and fat being measured from 6 months to 14 years of age [177]. Of the twelve studies evaluating BPA, eight found no statistically significant associations [113,172,173,174,176,177,178,179], one of 402 mother–child pairs in the USA reported a negative association with BMI, fat mass, and obesity risk at 2 to 9 years in girls [180]. Three other studies found positive associations with obesity markers. These three studies were conducted in Canada (N = 719) [181], China (N = 430) [182], and Republic of Korea (N = 788) [183]. Despite using a limited number of urine samples to assess the exposure (one or two), these studies reported positive associations between prenatal BPA and the waist-to-hip ratio at an average of 3.5 years (range 2–6 years), the waist circumference and an increased risk of obesity at 7 years, and the z-score weight from 2 to 6 years in girls. One study, from the SEPAGES cohort (N = 484), reported a positive association between second-trimester BPS and all infant growth parameters at 3 and 36 months of age [175]. No studies found an association between prenatal exposure to 2,4-dichlorophenol and 2,5-dichlorophenol. However, the number of studies on these compounds was relatively small (two to three). Among the small number of studies examining the association between BP-3 and childhood adiposity, a study of preadolescents (mean age = 11 years) in Spain (N = 1015) reported a positive association between prenatal exposure to BP-3 and a higher BMI z-score at 11 years [174].
Among the five studies of TCS, only one conducted in China reported a positive association with weight at 2 years using three spot urine samples (one from each trimester of pregnancy, N = 850 women–child pairs) [184]. This association appears to be stronger in girls. It is noteworthy that this study was one of the largest in sample size and number of urine samples collected during pregnancy.
We identified six studies regarding parabens. Four of them reported positive associations with postnatal weight, BMI, or the fat percentage at different ages, while the significantly associated paraben compounds differed between studies [172,178,185,186]. Three of these studies reported such effects in boys [178,185,186].
Regarding phthalates, 24 studies were identified. These studies assessed the association between prenatal phthalate concentrations and postnatal adiposity and growth. They were conducted in Australia, the USA, China, France, Germany, Greece, Mexico, Republic of Korea, and Spain. Sample sizes ranged from 180 [187] to 1301 [173]. Most of the studies measured phthalate metabolites in maternal urine collected at 1–3 time points during pregnancy. Serum was used in one study in Australia, resulting in lower detection frequencies and a higher risk of external contamination. In addition to these individual studies, a recent meta-analysis found a negative association between prenatal exposure to DEHP and child BMI [188]. A limitation of this meta-analysis was that only studies assessing associations with the molar sum of all the metabolites of DEHP were considered. This led to the exclusion of studies examining each metabolite separately, while several actually reported positive associations [79,189]. For other high-molecular-weight phthalates, isolated negative associations have been reported in boys between ∑DiNP and lean mass at 20 years [190], between ∑HMWP (high-molecular-weight phthalates) and weight gain during the first 6 months of life, and with BMI from 4 to 7 years [191]. A few studies also reported associations between exposure to MCPP, cx-MiNP, and MCOP and an increased risk of overweight [172,187], higher BMI [192], waist circumference [193], and lean mass [79]. For MBzP, the results were inconsistent, with four studies reporting positive associations [192,193,194,195] and two studies reporting negative associations [177,196] with obesity markers. Among the studies of low-molecular-weight phthalates, two studies reported positive associations between ∑LMWP (low-molecular-weight phthalates) and child BMI. However, this association was observed in boys in one study [190] and in girls in the other [79]. Regarding individual metabolites, of the 19 studies evaluating MEP, 6 studies reported positive associations with BMI [79,193,197], and only 1 study reported an inverse association with fat mass [187]. For MBP, a metabolite of DBP, most of the significant associations reported were positive, suggesting increased adiposity [193,198] and BMI [175,192,195]. Two studies have also reported positive associations between MiBP and BMI [79,190,192].
Overall, most studies examining the associations between prenatal phthalate exposure and child growth or BMI reported associations. However, the sign of the associations and the metabolites involved often differed. A meta-analysis, such as the one carried out for DEHP, could help to understand the links between prenatal exposure to phthalates and child growth. Although phthalates may act through common mechanisms, to date, very few studies have explored them as a mixture [79,113,172,173,199].
Table 3Effects of phenols, parabens, and phthalates on adiposity and growth of offspring using epidemiological data.
| Population | Class | Chemicals | Sampling Type, Number, and Timing | Timing of Outcome and Outcome of Interest | Model | Sex-Specific Analyses | Main Findings | Reference |
|---|---|---|---|---|---|---|---|---|
| INMA | Phenols and Phthalates | MEP, MBP, MiBP, MBzP, OH-MiNP, MECPP, MEHHP, MEOHP, MEHP, MCMHP, BPA | Urine | 7 years old | Uni-pollutant and multi-pollutant using PCA | Yes; p-values for interaction >0.2; | Compared to the first tertile, 7OHMMeOP was inversely associated with the BMI z-score in the third tertile and MECPP was inversely associated with overweight in both the second and the third tertiles of exposure. | [113] |
| CHAMACOS | Phenols, Phthalates, and Parabens | MEP, DEP, MBP, DBP, MiBP, DiBP, MBzP, BBP, DEHP, MEHP, MEHHP, MEOHP, MECPP, DiNP, MCOP, cx MiNP, DiDP, MCPP, MeP, PrP, TCS, 2,4-DCP, 2,5-DCP, BP-3, BPA | Urine | 5 years old | Uni-pollutant and multi-pollutant using BKMR | - | Urinary concentrations of MEP, cx-MiNP, and propylparaben were consistently associated with an increased BMI z-score and overweight/obesity status. | [172] |
| Raine Study | Phthalates | MEP, MiBP, | Serum | From birth and up to 20 years of age, longitudinal BMI z-scores, and DEXA at 20 years | Uni-pollutant, and linear mixed models with an interaction term between the phthalate level and age group (0–2, 2–11, and 11–20 years) | - | Compared to the lowest tertile: | [190] |
| EDEN | Phthalates | MEP, MBP, | Urine | Measures at birth and 5 years, prediction at 1, 3, and 5 years | Uni-pollutant | Only boys | No association between phthalate metabolite concentration and postnatal longitudinal weight. | [194] |
| ELEMENT | Phthalates | MEP, MBP, MiBP, MCPP, MBzP, MECPP, MEHHP, MEHP, MEOHP, ΣDEHP | Urine | Around 10 and 13 years old | Uni-pollutant, generalized estimating equation models with repeated-measures outcome | Yes, all analyses were sex-stratified | Among females, positive association between MBP T1, and BMI and MiBP T1, with the 3 outcomes. Negative association between MBzP T2 and skinfold thickness. | [195] |
| MIREC study | Phenols | BPA | Urine | Average 3.5 years old (range: 1.9–6.2) weight, height, waist/hip circumference, and subscapular/triceps skinfold thickness | Uni-pollutant, linear regression | Yes | BPA was positively associated with the waist-to-hip ratio among all children. | [181] |
| Mount Sinai | Phthalates | MEP, MBP, MiBP, MCPP, MBzP, ∑DEHP | Urine | Each follow-up visit scheduled at approximately ages 4–5.5 (mean, 4.9), 6 (mean, 6.2), and 7–9 (mean, 7.8) years old | Linear mixed-models with random intercepts to account for multiple observations per child; | Yes | No significant association between using continuous chemicals. In tertiles, compared with the lowest ∑DEHP tertile, the fat mass was 3.06% (95% CIs: –5.99, –0.09%) lower in the highest tertile. | [200] |
| The Mount Sinai Children’s Environmental Health Study (MSSM), | Phthalates | MEP, MBP, | Urine | At follow-up visits scheduled for approximately ages 4– 5.5, 6 and 7–9 years old (MSSM), 5 and 7 years old (CCCEH), and 4, 5, and 7–9 years old (HOME) | Bayesian modeling framework | Yes | MCPP was associated with increased odds of overweight/obese status overall. | [187] |
| Odense Child cohort | Parabens | MeP, EtP, | Urine | 7 years old | Uni-pollutant, linear regression | Yes, all analyses stratified by gender | Only n-BuP was positively associated with the total and android (visceral and subcutaneous) fat percentage among boys. | [185] |
| TIDES | Phthalates | MEP, MBP, | Urine | Birth, 1, 3, 4, and 6 years old | 1—Linear mixed-model with repeated growth measures; | Yes, differences by sex for weight models but not BMI models | MEP, MBzP, MiBP, and MBP were all inversely associated with weight and BMI at birth. | [192] |
| Ma’anshan Birth Cohort (MABC) | Phthalates | MMP, MEP, | Urine | Birth, 3, 6, and 9 months, and 1, 1.5, 2, 2.5, 3, 4, 5, and 6 years old | Uni-pollutant (linear/logistic regression and GEE with repeated measurement) and multi-pollutant (quantile g-computation and BKMR) | Only girls | Pregnancy means ΣLMW and MBzP were negatively associated with BMI z-score at birth. | [201] |
| INMA | Phenols, Parabens, and Phtalates | MEP, MBP, MBzP, DEHP, MeP, EtP, PrP, n-BuP, BP-3, BPA | Urine | 11 years old | Uni-pollutants (generalized additive mixed models); multi-pollutant (BKMR) | Yes | BP-3 was associated with a higher BMI. | [174] |
| HELIX project | 45 ED compounds | Among them, 7 high-molecular-weight | Urine and blood samples during pregnancy | 6 to 11 years old | Associations were assessed using Bayesian weighted quantile sum regressions applied to mixtures for each chemical | Yes | ΣLMW and ΣHMW of phthalates: ↓ MetS risk score. | [202] |
| China | Parabens | MeP, EtP, PrP, n-BuP, BzP | Urine | 3 years old | Uni-pollutant generalized linear models | Yes | Among all and among boys: urinary EtP concentrations were positively associated with weight z-scores. | [186] |
| Sheyang Mini Birth Cohort Study | Phenols | BPA | Urine | 7 years old | Uni-pollutant, generalized linear regression models, and multivariable logistic regression models | Yes, no significant interaction | Positive associations with the waist circumference. | [182] |
| CHAMACOS | Phthalates | MEP, MBP, MiBP, MCPP, MCOP, cx-MiNP, MBzP, ∑DEHP | Urine | 5, 7, 9, 10.5, and 12 years old | Uni-pollutant, GEE with repeated outcomes | Yes | MEP, MBzP, and ΣDEHP were associated with a 20–30% increase in the odds of overweight/obesity at each age point in both boys and girls. MBP was associated with a 30–40% increase in odds in boys only. | [193] |
| CHAMACOS | Phenols | BPA | Urine | 2, 3.5, 5, 7, and 9 years | Uni-pollutant, association with outcomes at 9 years: linear/logistic regressions. | Yes | Association was not significant among all, while there was a negative association among girls (significant negative association with BMI, body fat, and risk of overweight/obesity), and null associations among boys. | [180] |
| CHAMACOS | Phthalates | MBP, MEP, MiBP, MBzP, cx-MiNP, MCOP, MCPP, MEHP, MEHHP, MECPP, MEOHP | Urine | 11 follow-up visits between 2 and 14 years old | PCA | Yes | MEP-positive association with BMI through 12 years. | [197] |
| HOME | Phenols | TCS | Urine | 8 years old | Uni-pollutant, linear regression | yes, all p-values = 0.37 | No significant association. | [203] |
| PROGRESS | Phthalates | ΣDEHP, ΣDiBP, ΣDiNP, ΣDBP, MBzP, MECPTP, cx-MiNP, MCPP, MEP | Urine | 4, 6 and 8 years old | Uni-pollutant (linear and multinomial logistic regressions) and mixture (quantile G-computation) | Yes | In comparison to the “low-stable” group, | [199] |
| EDC Cohort | Phthalates | MEHHP, MEOHP, | Urine | 6 years old | Uni-pollutant, linear regressions | Yes | MEHHP and MnBP were inversely associated with the BMI z-score among girls. | [204] |
| MOCEH Study | Phenols | BPA | Urine | Birth, and 6, 12, 24, 36, 60, and 72 months | Uni-pollutant, linear regression for each outcome, and LMM for growth from 6 to 72 months | Yes | BPA was positively associated with the weight z-score at birth among all children and among boys, while positively associated among girls at 24, 36, 60, and 72 months. | [183] |
| Human cohort conducted at Wuhan, | Phthalates | MECPP, | Urine | Birth, 6, 12, and 24 months old | Uni-pollutant—mixed linear models | Yes | MECPP and MEOHP were positively related to the average weight z-scores in male offspring. | [189] |
| CCCEH | Phthalates | MEHP, MEHHP, | Urine | 5 (only BMI) and 7 years old | Analyses by DEHP component and non-DEHP component; | Yes | Non-DEHP components were negatively associated with all outcomes of interest among boys. | [205] |
| German LIFE Child cohort study, | Phthalates | MBzP, MEHP, MEHHP, MEOHP, MECPP, MCIOP, oxo-MiNP, OH-MiNP, | Urine | Birth, 1 and 2 years old | Uni-pollutant, robust linear regression | Yes | ∑HMW was negatively associated with birth weight in all children and among girls (not boys). | [206] |
| Shanghai Obesity and Allergy Birth Cohort Study, | Phenols | BPA | Urine | 2 years old | Uni-pollutant, linear regression | Yes, all analyses stratified by gender | No differences were found in any adiposity measures. | [176] |
| EDEN | Phenols and Parabens | 2,4-DCP, 2,5-DCP, | Urine | 6 months, 1, 2, and 3 years old | Uni-pollutant | Only boys | Parabens were positively associated with weight at birth; remained for 3 years for methylparaben. | [178] |
| HOME, | Phthalates | ∑DEHP, | Urine | 8 years | Uni-pollutant, 2 samples averaged for prenatal exposure estimation | Yes | MBzP was negatively associated with the BMI z-score. | [196] |
| Rhea Study, | Phthalates | MEP, MBP, | Urine | 4 and 6 years | Uni-pollutant, GAMs to explore shape GEE for associations with repeated outcomes | Yes | Only MnBP was associated with a change in waist-to-height ratio at ages from 4 to 6 years old. | [198] |
| Rhea Study, | Phenols | BPA | Urine | Birth (weight), from 6 to 4 years old (BLI, at 4 years for other outcomes) | Uni-pollutant, first GAM to explore the shape of the relationships, then linear regressions; mixed effects linear regression model for BMI trajectories | Yes | No significant association, but patterns of BMI trajectories differed between boys and girls (only visually; not tested). | [179] |
| INMA-Sabadell | Phthalates | ΣHMW, | Urine | From birth to 6 months old for weight gain, and 1, 4, and 7 years old for BMI | Uni-pollutant and one multi-pollutant model simultaneously adjusted for ΣHMW and ΣLMW; | Yes | ΣHMW was inversely associated with the weight gain z-score (from 0 to 6 months) and BMI z-scores at 4 and 7 years old among boys, while associations tended to be positive among girls. Zero associations among all children. | [191] |
| HELIX | Phenols, Parabens, and Phthalates | MEP, MiBP, | Urine | One visit between 6 and 11 years old | Uni-pollutant and DSA | Not with phenols and phthalates | No significant association. | [173] |
| China | Phenols | TCS | Urine | 6 months, 1 and 2 years old | Uni-pollutant | Yes | Prenatal exposure (mean expo) to triclosan was associated with elevated 2-year-old weight z-score in all and in girls. | [184] |
| ELEMENT | Phenols and Phthalates | BPA, MBP, MBzP, MCPP, MEP, MiBP, MEHP, MECPP, MEHHP, MEOHP, ΣDEHP, ΣHMW, ΣLMW | Urine | One single visit between 8 and 14 years old | Uni-pollutant, linear regression | Yes, all p-values of sex interaction > 0.2 | Only MBzP was negatively associated with the BMI z-score. | [177] |
| SEPAGES cohort | Phenols, Parabens, and Phthalates | 4 phenols, 4 parabens, 7 phthalates, and 1 non-phthalate plasticizer | Weekly pooled urine samples collected from the mother during T2 and T3 trimesters | BMI, height, weight, and head circumference at 3 and 36 months of age | Associations with individual chemicals using adjusted linear regression and mixtures of chemicals using a Bayesian kernel machine regression model | Yes | BP at T2: ↓ BMI and weight at 36 months of age. | [175] |
BBP: benzyl butyl phthalate; n-BuP: n-butylparaben; BP-3: benzophenone-3; BPA: bisphenol A; BzP: benzylparaben; DBP: di-n-butyl phthalate; DEHP: di-(2-ethylhexyl) phthalate; DEP: diethyl phthalate; DiBP: diisobutyl phthalate; DiDP: di-isodecyl phthalate; DiNP: di-isononyl phthalate; EtP: ethylparaben; HMW: high-molecular-weight phthalates; LMW: low-molecular-weight phthalates; MBP: mono-n-butyl phthalate; MBzP: monobenzyl phthalate; MCIOP: mono-(4-methyl-7-carboxyheptyl) phthalate; MCMHP: mono [2-(carboxymethyl)hexyl] phthalate; MCOP: mono-carboxy-iso-octyl phthalate; MCPP: mono(3-carboxypropyl) phthalate; MECPP: mono-(2-ethyl-5-carboxypentyl) phthalate; MECPTP: mono-2-ethyl-5-carboxypentyl terephthalate; MEHHP: mono(2-ethyl-5-hydroxyhexyl) phthalate; MEHP: mono(2-ethyl-hexyl) phthalate; MEOHP: (mono(2-ethyl-5-oxohexyl) phthalate; MEP: mono-ethyl phthalate; MetS: risk of metabolic syndrome; MiBP: mono-iso-butyl phthalate; MiDP: mono-iso-decyl phthalate; MiNP: mono-iso-nonyl phthalate; cx-MINP: mono-carboxy-iso-nonyl phthalate; MMP: monomethyl phthalate; MNP: mono-isononyl phthalate; MeP: methylparaben; OH-MiNP: mono-hydroxy-iso-nonyl phthalate; oxo-MiNP: mono-oxo-iso-nonyl phthalate; PrP: propylparaben. Arrows indicate the direction of the effect: a downward arrow (↓) indicates a decrease and an upward arrow (↑) indicates an increase.
6.1.2. Maternal Exposure to PPPs and Metabolic Outcomes in Experimental Models (Table 4)
Studies of exposure to PPPs were initially interested in the risk of allergies in offspring, and also on the risk of mammary and genital carcinogenesis [207]. These will not be addressed in this review. These studies then focused either on the reproductive toxicity for the offspring, discussed later, or on the neurodevelopmental effects and their repercussions on behavior and learning. Both of these subjects were reported in a very recent review (for the bisphenols and phthalates) [208]. Finally, more recently, studies have focused on the metabolic effects of these PPP exposures and the repercussions on bodyweight, which will be the subject developed here.
Maternal oral exposure to bisphenol during the perinatal period, either A, S, or F, resulted in dose-grading adverse effects in female offspring. These adverse effects on the intestinal and systemic immune response depend on the bisphenol nature, i.e., A, S, or F. In mice, stronger impacts on inflammatory markers in feces were observed with BPS at the dose of 0.005 mg/kg bw/d. Exposure to BPA and BPF at low doses induced significant immune response changes in the offspring. These changes led to both intestinal and systemic Th1/Th17 inflammation [209]. Additionally, maternal exposure to BPS resulted in adverse effects on the triacylglycerol (increase in males), hormone levels (increase in T3 males and T4 females), and behavior (decreased food intake) of the offspring. These effects were observed as a function of the dose with a non-monotonic response [210]. Maternal exposure to BPA induced lighter weaning weights in some males. These males experienced rapid catch-up growth immediately after weaning [211]. So, they represent a subpopulation sensitive and vulnerable to very low fetal serum concentrations of BPA in the pg/mL range, particularly in the case of glucose intolerance. Gestational exposure to BPA has been shown to upregulate offspring pancreatic β-cell division and mass in an ERβ-dependent manner in adult male mice [212]. In C57BL/6J, maternal exposure to BPA mimicking human exposure levels (from 0.010 to 10 mg/kg bw/day) led to dose-specific effects on pancreatic islets in both the first (F1) and second generations (F2) in males only. Moreover, an increase in bodyweight was observed only in the F3 males. In addition, the lowest dose reduced the β-cell mass and smaller islets associated with increased insulin secretion, without a change in glucose tolerance. However, changes in the cytokine levels were reported across the generations in males [213]. In a transgenerational context, female offspring from males exposed in utero and during lactation to BPA were shown to exhibit impaired glucose tolerance despite the absence of compromised insulin sensitivity in vivo or reduced ex vivo glucose-stimulated insulin secretion. However, male offspring showed normal glucose tolerance [214]. Prenatal exposure to BPA (5 mg/kg bw/day) disrupted hepatic lipid homeostasis in a sex- and age-dependent manner. These effects, investigated by lipidomic and transcriptomic approaches, were marked around weaning (the accumulation of lipids and inflammation of the liver), but tended to fade with age, especially in females (bodyweight and total lipid content) [215,216].
In sheep, prenatal exposure to BPA (0–0.05–0.5–5 mg/kg bw/day by subcutaneous injections from days GD30 to GD90; term: 147 days) induced peripheral insulin resistance and adipose tissue disruptions in female offspring of 21 months of age [217]. In addition, lipotoxicity (an increase in blood and tissue triglycerides), accompanied by an increase in oxidative stress, was observed in a non-monotonic manner. In parallel, a reduction in antioxidants was shown in both liver and skeletal muscles, as well as altered proinflammatory markers in the liver (an increase in TNF-α and CD68, but IL-6 and IL-1B) and skeletal muscle (an increase in IL-6, IL-1B, CCL2, and CD38) [218]. These effects contributed to the resistance to insulin. These BPA-induced prenatal metabolic dysfunctions were corroborated for the middle dose (0.5 mg/kg/day) by transcriptomic analyses. These data provided mechanistic clues to explain oxidative stress and lipid accumulation, and potential mitochondrial and fibrotic defects in these tissues [219]. In the same animal model, prenatal exposure to BPA (0.05, 0.5, or 5 mg/kg/day) by subcutaneous injections from days GD30 to GD90 induced a trend toward decreased insulin and β-cell counts associated with an increase in glucagon and α-cell counts. The results were most consistent at the lowest BPA dose in fetal pancreata at GD90 and in adult offspring. These data suggest that early-life BPA exposure poses a likely threat to metabolic health [220].
In mice, maternal subcutaneous exposure to BPA (0.010 or 0.100 mg/kg bw/day) from GD9 to GD16 impaired pancreatic function, with a decrease in the mass of β-cell [221] and impaired glucose homeostasis in 6-month-old male offspring, but not in females [222]. The same replicated protocol showed that such maternal exposure of PND30 increased the retinoid concentrations and gene expression of key elements involved in the retinoid system in the liver in male offspring [223].
In mice, maternal exposure to BPA throughout gestation and lactation (1000 nM via drinking water) increased the cytokine levels in the spleens of the PND21 and PND42 offspring. The levels of cytokines were derived from Th17 cells (IL-17 and IL-21) and were essential for the differentiation of Th17 cells (IL-6 and IL-23). These increases were more pronounced in females and started at the lowest dose of BPA (100 nM), promoting inflammation induced in offspring [224]. When the mice were orally exposed to BPA (0.000050, 0.050, or 50 mg/kg diet) throughout gestation and lactation, primary bone marrow-derived mast cells (BMMCs) presented, after activation, increased cysteinyl leukotriene and TNFα production in all of the exposed groups [225]. Additionally, an increase in prostaglandin D2 and IL-13 production was observed only in the most BPA-exposed group. These BMMCs were generated from the bone marrow culture extracted from the femurs of PND21 progeny [225]. This production of proinflammatory mediators is generally associated with asthma, another harmful effect linked to the exposure of BPA. In mice exposed to BPA (0.010 or 10 mg/kg bw/d) throughout gestation and lactation, males exposed to the lowest dose were lighter at birth. These males then underwent rapid catch-up growth until weaning and showed increased bodyweight after PND117. Furthermore, glucose intolerance and insulin resistance resulted in increased body fat in offspring exposed to the higher dose [226]. Another study exposing the dam to BPA (0–30 mg/kg bw/day) for the same periods as above showed dose-dependent increases in body and liver weights in 20-week-old offspring [227]. A dose-dependent decrease in circulating glucagon in male offspring was also observed. However, in this study, females showed a dose-dependent decrease in body, liver, muscle, and fat pad weights. The latest effects in females were accompanied by decreased adipocyte size, serum lipids, serum leptin, and adiponectin, in parallel with increased physical activity [227]. Maternal exposure to BPA through drinking water (10 mg/L) induced insulitis, i.e., a pancreatic disease caused by lymphocyte infiltration, and accelerated the prevalence of diabetes in the 20-week-old female offspring [228]. In mice, perinatal exposure to BPA via the maternal diet (0.000050, 0.050, or 50 mg BPA/kg diet) induced an increase in the energy expenditure of the offspring during life until the age of 10 months. In female offspring, BPA exposure tended to decrease bodyweight and fats, with improved glucose and insulin profiles, at the highest dose, leading to a hyperactive and lean phenotype [229].
Maternal exposure of mice to the antibacterial TCS (8 mg/kg bw/d) from GD6 to GD14 has been shown to alter prenatal and postnatal growth and development, as well as metabolic phenotypes in male and female offspring. Compared with control offspring, TCS offspring (male or female) initially showed reduced bodyweight at birth, but then showed more rapid bodyweight gain during the fifth day of gestation, which increased over time [160]. Indeed, PND30 overweight TCS offspring showed, at PND60, increased visceral fat and adipocyte size, with delayed glucose clearance and insulin resistance. In rats, maternal oral exposure to TCS (1 mg/kg/day from GD14 to PND20 during lactation) increased bodyweight, blood glucose, and cholesterol in 5-month-old offspring, as well as food intake [230]. The increase in the latter is probably due to the increased hypothalamic expression of orexigenic neuropeptides. Moreover, in aged rats with high in utero exposure to TCS, a decreased hepatic glycogen content was observed, while the serum and hepatic triglyceride content increased with the upregulation of genes involved in carbohydrate and lipid metabolism pathways in the liver [231]. This was also observed in mice, along with an increase in serum and liver triglycerides with the increased gene expression of a protein involved in fatty acid synthesis, but the decreased gene expression of a protein involved in fatty acid oxidation [232].
What has been shown for any of these pollutants, depending on the dose studied, the route of exposure, the group of offspring considered (male or female), their age at the time of analysis, in addition to the animal model and species, must be rethought when it comes to combining these pollutants, as all of the cards of effects have to be reshuffled. The study of cocktail effects is likely to yield surprising results, like those investigated in a few studies [233], with a reduction in live weight for all of the mixtures in females regardless of the dose, and in males only at mid-dose.
Maternal n-BuP (parabens) exposure in mice was shown to induce a higher food intake and weight gain in female offspring only, probably reducing hypothalamic POMC expression induced by epigenetic modification [234].
According to the literature, early-life exposure to DEHP is potentially associated with increased adiposity in rodents [235,236]. Perinatal exposure to different phthalates will cause different metabolic outcomes in mice, with sex-specific responses. In mice, females exposed in utero to DEHP exhibited increased body fat and decreased lean mass, whereas exposure to DiNP induced only a decrease in glucose tolerance. In contrast, prenatal exposure of males to phthalate did not lead to any significant differences in the measured metabolic outcomes [237]. However, other studies in mice have reported that prenatal exposure to DEHP at a low dose (0.2 mg/kg/day) led to a metabolic syndrome in male offspring, including abnormal adipogenesis, energy expenditure, and glucose metabolism, by the deregulation of hepatic thiamine transport enzymes [238]. In the mouse liver, perinatal exposures to phthalates were associated with the short- and long-term activation of PPAR target genes, which was manifested by increased fatty acid production in early postnatal life and increased fatty acid oxidation in adulthood [239]. A similar accumulation of hepatic lipids was observed in rats exposed to DEHP during the perinatal period [240], as described for TCS [231,241]. Additionally, maternal exposure to DEHP in rats (0.75 mg/kg bw/day from GD6 to PND21) decreased serum insulin and triglyceride levels in PND70 male offspring. These outcomes were linked by the authors to the elevated expression of PPARγ (mRNA and protein) in white adipose tissue [242]. At PND21, male and female offspring pups exposed in utero to DEHP (700 mg/kg bw/day during the last third of gestation and lactation) and DBP (500 mg/kg bw/day) showed increased fasting glucose levels, as well as metabolic alterations [243]. Indeed, gestational exposure to DEHP has been shown to promote β-cell dysfunction and whole-body glucometabolic abnormalities in F1 offspring by downregulating the expression of critical genes involved in β-cell development and function [244]. Maternal exposure to a lower dose of DEHP (10 and 100 mg/kg bw/day) from GD9 to PND21 (lactation period) by oral gavage in the male offspring induced hyperglycemia, impaired tolerance to glucose and insulin, as well as hyperinsulinemia at PND80. This phenotype occurs because the levels of insulin signaling molecules such as insulin receptors, IRS1, Akt, and its phosphorylated forms are reduced [245]. In addition, maternal exposure to DEHP throughout gestation has been shown to disrupt thyroid function in offspring pups. DEHP disrupts thyroid function by damaging thyroid follicles and affecting thyroid transcription factor 1 (TTF-1), paired box 8 (PAX8), sodium iodide symporter (NIS), and thyroid peroxidase (TPO) both at the transcriptional level and at the protein level. In pups, this damage leads to a reduction in total thyroxine (T4) and an increase in thyroid-stimulating hormone (TSH) [246]. In adults, this damage was observed without an altered macro-index such as bodyweight in males at 14 weeks. The males were then likely to develop insulin resistance (hyperinsulinemia), oxidative stress (increased CAT catalase), and hypothyroidism (decreased T4) [247].
Prenatal exposure to low doses of DEHP or other phthalates has resulted in life-long metabolic consequences in a sex-dependent manner in offspring. Most of the time, the results are contradictory, with effects only in females or males. This exposure suggests a potential risk factor for later obesity and metabolic syndrome development in adulthood. Additionally, most studies on perinatal exposure to phthalates introduce bias to the extent that they focus only on the effects on male offspring in rodents, and only on female offspring in sheep.
As we can see in Table 4, the experiments relating to maternal oral exposure to phenols, parabens, and/or phthalates during the perinatal period (all or part of gestation and/or lactation) have studied very varied doses of the order of mg/kg with non-monotonic dose–responses and sex-specific metabolic effects in offspring. The observed effects are often transitory around weaning and deserve to be studied later in adulthood.
Table 4Synthesis of the offspring outcomes related to maternal oral exposure to phenols, phthalates, and/or parabens during the perinatal period (all or part of gestation and/or lactation) according to the animal model, the chemicals, the dose administrated, the exposure route and duration.
| Animal Model | Chemicals | Dose Administered | Exposure Route | Exposure Duration | Observation Stage | Function Studied | Metabolic Postnatal Outcomes | Additional Outcomes | Reference |
|---|---|---|---|---|---|---|---|---|---|
| Mouse | BPA | 0.01, 10 mg/kg bw/d | Oral route (food) | Paternal exposure (12 weeks from 5 weeks of age) | 0–20 weeks F1 | Metabolism | Glucose tolerance and body composition: no effect | Insulin-dependent glucose disposal in post-pubertal: BPA ↑ (0.01) | [214] |
| Mouse | BPA | 500 mg/kg bw/d | Oral administration | GD8-GD14 | PND56 | Reproduction/fertility | Mortality at birth: BPA ↑ | Number of mature spermatozoids: BPA ↓ (♂) | [248] |
| Mouse | BPA | 0.01, 0.1 mg/kg bw/d | Subcutaneous injections | GD9-16 | PND1-21 + 6 months | Glucose homeostasis | Bodyweight at birth and PND21: BPA ↑ (0.01)/↓ (0.1) | Litter size: no effect Maternal glucose intolerance and total mean area under the curve of glucose tolerance: BPA ↑ (0.01) | [222] |
| Mouse | BPA | 5 mg/kg bw/d | Oral gavage | GD1-20 | PND5-3 weeks of age | Metabolism | Bodyweight: BPA ↓ | Number of differentially expressed genes (DEGs): 86 (adipose tissue)/93 (hypothalamus) | [249] |
| Mouse | BPA | 0.01, 10 mg/kg bw/d | Oral route (food) | GD-14 to PND21 | PND7, 14, and 21 weeks of age | Metabolism | F2 and F3 bodyweight: no effect | Proinflammatory cytokine levels: BPA ↓ (IL-1β and IL-12p70) | [213] |
| Mouse | BPA | 0.1, 1, or 10 mg/L | Oral route (drinking water) | GD0-PND21 | 7 to 28 weeks of age | Metabolism | Mean insulitis grade and number of diabetic mice: BPA ↑ (10) | Activated caspase-3-positive, insulin-positive, and glucagon-positive apoptotic cells in pancreatic islets: BPA ↑ | [228] |
| Mouse | BPA | 0.01 mg/kg bw/d | Subcutaneous injections | GD9-16 | PND0 and 30 | Pancreatic function | Pancreatic β-cell area, mass, and proliferation: BPA ↑ (♀) | [212] | |
| Mouse | BPA | 0.01, 0.1 mg/kg bw/d | Subcutaneous injections | GD9-16 | PND30 | Retinoid signaling pathway | ATRA (all-trans-retinoic-acid) hepatic concentration: BPA ↑ (0.01) | [223] | |
| Mouse | BPA | 0.01, 0.1 mg/kg bw/d | Subcutaneous injections | GD9-16 | PND0, 21, 30, and 120 | Pancreatic function | Bodyweight: BPA ↓ (PND0/PND21: 0.1)/↑ (PND21-30) | DEGs involvement: cell cycle/mitosis/cell division | [221] |
| Mouse | BPA | 10, 100, or 1000 nM | Oral route (drinking water) | GD0-PND21 | PND21, 42 | Spleen immune response | Litter size, birth weight, survival rate, and sex ratio: no effect | Gestational weight: no effect | [224] |
| Mouse | BPA | 0.00005, 0.05, or 50 mg/kg food/d | Oral route (food) | GD-14 to PND21 | 6 months of age | Inflammatory mediators/asthma | Bone marrow-derived mass cell (BMMC) cysteinyl leukotriene secretion: BPA ↑ | - | [225] |
| Mouse | BPA | 0.01, 10 mg/kg bw/d | Oral route (food) | GD-14 to PND21 | GD9.5-10.5, GD16.5-17.5, PND1, 14, 21, 28, and 98-117 | Metabolism | Fetal bodyweight at PND1, 14, and 21: BPA ↓ (♂: 0.01) | F2 total insulin-like growth factor 2 (Igf2) mRNA expression: BPA ↑ (10) | [226] |
| Mouse | BPA | 0.006, 0.06, or 0.6 mg/kg bw/d | Subcutaneous capsule | GD9 to at least 3 weeks | PND2, 21, and 3 to 12 weeks of age | Glucose metabolism | Bodyweight at weaning and 5 weeks: BPA ↓ (♂: 0.006) | [211] | |
| Mouse | BPA | 0.003, 0.01, 0.03, 0.1, 0.3, 1, or 3 mg/kg bw/d | Oral route (food) | GD-14 to PND21 | PND21, and 5 to 23 weeks of age | Metabolism | Litter size, sex ratio, and survival rate: no effect | Mortality: no effect | [227] |
| Mouse | BPA | 0.00005, 0.05, or 50 mg/kg diet | Oral route (food) | GD-14 to PND21 | 3, 6, 9, and 10 months of age | Metabolism | Litter size, survival rate and sex ratio: no effect | Oxygen consumption: BPA ↑ (♀ 3 months: 50/♀ 6 months: 0.05, 50/♀ 9 months: 0.00005/♂ 9 months: 0.05, 50) | [229] |
| Rat | BPA | 0.05 or 5 mg/kg bw/d | Oral gavage | GD3 to GD18 | PND21 and PND60 | Lipid metabolism | Total lipid content: BPA ↓ (♀ PND21: 5) | PUFAs: BPA ↑ (PND21) | [215] |
| Rat | BPA | 0.05 or 5 mg/kg bw/d | Oral gavage | GD3-18 | PND1, 13, 21, 30, and 60 | Liver metabolism | Bodyweight: BPA ↑ (♀ PDN1-PND60/♂ PND13, PND21-PND30/♂ PND60: 5) | Anogenital distance (AGD): BPA ↑ | [216] |
| Rat | BPA | 0, 0.05, 0.5, or 5 mg/kg bw/d | Oral gavage | GD5-19 | PND1, 21 and 56 | Lipid metabolism | Bodyweight: BPA ↓ (♀ PND56: 0.5) | Liver mTOR mRNA expression: BPA ↑ (PND21: 0.05/PND56: 5) | [250] |
| Sheep | BPA | 0.05, 0.5, or 5 mg/kg bw/d | Subcutaneous injections | GD30-90 | GD68, 6, 14, 15, 19, 21 weeks, and 13 months of age | Metabolism | Fasting glucose level: BPA ↑ (6 weeks: 0.05) | [217] | |
| Sheep | BPA | 0.5 mg/kg bw/d | Subcutaneous injections | GD30 to GD100 | 21 months F1 | Liver and muscle metabolic function | Number of differentially expressed (DEGs) genes in the liver: BPA ↓ (138)/↑ (56) | Number of differentially expressed genes in the muscle: BPA ↑ (80)/↓ (32) | [219] |
| Sheep | BPA | 0.05, 0.5, or 5 mg/kg bw/d | Subcutaneous injections | GD30-90 | 21 months F1 | Metabolism | Liver IL-1β, IL-6, and chemokine (C-C) ligand 2 expression: BPA ↓ | Skeletal muscle IL-6 expression: BPA ↑ | [218] |
| Sheep | BPA | 0.5 mg/kg bw/d | Subcutaneous injections | GD30 to GD100 | 21 months F1 | Cardiovascular function | Lung, kidney, and adrenal weight: BPA ↓ | Cardiac failure gene expression in the left ventricle: BPA ↑ (ANP, COL1A1) | [251] |
| Sheep | BPA | 0.05, 0.5, or 5 mg/kg bw/day | Subcutaneous injections | GD30-GD90 | GD65, GD90, 21 months F1 | Pancreas | Fetal weight: no effect (GD65, GD90) | [220] | |
| Rat | BPS | 0.01, 0.05 mg/kg bw/d | Oral gavage | GD1-PND21 | PND21, PND160-180 | Metabolism, endocrine system | Plasma triacylglycerol and thyroxine (T4) level: BPS ↓ (♂ PND21: 0.05) | Plasma testosterone level: BPS ↓ (0.05) | [210] |
| Mouse | BPA, BPS, BPF | 0.005, 0.05 mg/kg bw/d | Oral route | GD15-PND21 | PND70 | Intestinal and systemic immune response | Bodyweight: BPF ↓ (0.05: PND70) | Plasma-specific anti-E. coli IgG: BPS ↑ (0.05) | [209] |
| Mouse | BPA + NP, OP or IsoBP | 5, 50, or 500 mg/kg bw/d | Oral gavage | GD1-21 | PND1, and 21 and 41 | Fetal development | Survival rate: BPA + OP ↓ (5) | Dams gestational day times: BPA + IsoBP ↓/BPA+ NP ↓ (50, 500)/BPA + OP (50, 500) | [233] |
| Rat | BPA, DEHP | 5 µg/kg bw/d (BPA) and/or 5µg and 7.5 mg/kg bw/d (DEHP) | Oral administration | GD6-GD21 | PND1, 7, 14, and 21, PND21 to PND84, PND112, and PND168 | Fetal development/food intake | Maternal gestational weight gain: no effect | Abortion rate: BPA + DEHP ↑ | [158] |
| Mouse | TCS | 4, 8 mg/kg bw/d | Oral gavage | GD6-14 | PND1, 30, and 60 | Placenta/metabolism | Bodyweight gain: TCS ↑ (PND5, PND30, PND60: 8) | [160] | |
| Rat | TCS | 1 mg/kg food | Oral route (food) | GD14-PND20 | PND1 to 8 months | Food intake/metabolism | Bodyweight: TCS ↓ (♂ 2 months)/↑ (♂ 4 months/♀ 8 months) | Npy and Agrp hypothalami mRNA expression (involved in the appetite regulatory network): TCS ↑ (♂) | [230] |
| Rat | TCS | 10, 50 mg/kg bw/d | Oral gavage | GD0-PND21 | 3, 21, and 52 offspring’s weeks of age | Metabolism/microbiota | Glucose clearance: TCS ↓ (52 weeks: 50) | Alpha diversity of the microbial community in gut microbiota: TCS ↓ (50) | [231] |
| Mouse | DEHP | 10–12, 55–64, and 119–145 mg/kg bw/d | Oral route (food) | GD18 | Fatty acid metabolism | α-linolenic acid level: DEHP ↑ (10–12) | Linoleic and oleic acid level: DEHP ↓ (55–64, 119–145) | [252] | |
| Mouse | DEHP | 0.2, 2, or 20 mg/kg bw/d | Oral gavage | GD-7 to GD21 | 12 weeks of age | Metabolism | Bodyweight at 12 weeks of age: DEHP ↑ (♂: 0.2) | Expression of thermogenic genes: DEHP ↓ (0.2 (Ucp1, Cidea, Adrb3)) | [238] |
| Mouse | DEHP | 25 mg/kg chow/d (5 mg/kg bw/d) | Oral route (food) | GD-14 to PND21 | 5 months of age | Cardiac DNA methylome | Mortality, litter size, sex ratio, bodyweight, and relative heart weight ratio: no effect | [253] | |
| Mouse | DEHP | 250 mg/kg bw/d | Oral gavage | GD0-PND21 | PND21 | Skeletal muscle development | Bodyweight change: DEHP ↓ | [254] | |
| Mousse | DEHP | 30 mg/kg bw/d | Oral administration | GD-28 to PND28 | 8 weeks of age | Metabolism | Diastolic arterial, systolic, and mean blood pressure: DEHP ↑ | Aorta phosphorylated eNOSthr497: DEHP ↑ | [236] |
| Rat | DEHP | 300, 600 mg/kg bw/d | Oral gavage | PND21 to 24 weeks (males) | 25 weeks | Metabolism | Accumulation of DEHP in fat and serum: DEHP ↑ (long-term) | [255] | |
| Rat | DEHP | 0.75 mg/kg bw/d | Oral gavage | GD6-PND70 | PND0-70 | Lipid metabolism | Bodyweight: DEHP ↑ (GD6-PND21)/↓ (GD6-PND70) | [240] | |
| Rat | DEHP | 300 mg/kg bw/d | Oral gavage | GD14-PND0 | PND60, 200 | Blood pressure/heart rate | Systolic blood pressure: DEHP ↓ (nighttime) | Locomotor activity during daytime: DEHP ↓ (PND60) | [256] |
| Rat | DEHP | 20, 50, 100, 300, and 750 mg/kg bw/d | Oral gavage | GD14-PND0 | PND60 | Metabolism | Circulating aldosterone concentration: DEHP ↓ (100, 300, 750)/↑ (♀: 300) | Circulating testosterone concentration: DEHP ↓ (100, 300, 750) | [257] |
| Rat | DEHP | 0.75 mg/kg bw/d | Oral administration | GD6-PND21 or/and PND22-70 | PND70 | Metabolism | Serum insulin and TG levels: DEHP ↓ (GD6-PND21) | [242] | |
| Rat | DEHP | 10 or 100 mg/kg bw/d | Oral gavage | GD9-PND21 | PND80 | Glucose metabolism | Fasting blood glucose level and insulin resistance: DEHP ↑ | Serum testosterone and estradiol concentration: DEHP ↓ | [245] |
| Rat | DEHP | 1, 10, or 100 mg/kg bw/d | Oral gavage | GD9-21 | PND60 | Pancreatic function | Fasting blood glucose level: DEHP ↑ | Islets mRNA expression: DEHP ↑ (FoxO1, Atf4, Atf6, Bip, DNMT3a, Cnmt3b, MeCP2, MDB2, mir143, and mir375/10, 100: DNMT1)/↓ (10, 100: GLUT2) | [244] |
| Rat | DEHP | 600 mg/kg bw/d | Intragastrically administration | GD0-PND21 | 8 and 14 weeks of age | Metabolism | Body and organ weight, body length, food intake, fat pad weight, and ratio: no effect | Serum catalase level: ↓ | [247] |
| Rat | DEHP | 1, 10, or 100 mg/kg bw/d | Oral gavage | LD1-21, lactational | PND22 | Cardiac metabolism/function | Fetal bodyweight: DEHP ↓ (PND9-22) | Heart weight and deoxyglucose uptake: DEHP ↓ (PND22) | [258] |
| Rat | DEHP | 300 mg/kg bw/d | Oral gavage | 12 weeks, postnatal (adult) | adult | Liver metabolism | Bodyweight: no effect | Liver malondialdehyde content: DEHP ↑ | [241] |
| Rat | DEHP | 30, 300, or 750 mg/kg bw/d | Oral gavage | GD0-PND21 | PND7, 14, and 21 | Thyroid function | Serum total T3 level: no effect | PND14 deiodinases (Dio) mRNA expression: DEHP ↑ (30: Dio2/300: Dio1, Dio2/750: Dio1, Dio2)/↓ (Dio3) | [246] |
| Rat | DEHP | 1, 20, 50, 100, or 300 mg/kg bw/d | Oral gavage | GD14-PND0 | GD19, PND3, and 21 and 60 | Metabolism | PND21 number of adrenal glands differentially expressed genes: DEHP ↑ (100 (67), 300 (175)/↓ (100 (238), 300 (252)) | Immune response pathways: DEHP ↑ | [259] |
| Rat | DEHP | 0.25, 6.25 mg/kg bw/d | Oral gavage | GD0-PND21 | PND0, 2, 15, and 21 | Kidney development and fetal metabolism | Birth, PND15-21 bodyweight: DEHP ↓ (6.25) | Maternal postpartum weight, litter size, and sex ratio: no effect | [260] |
| Rat | DBP | 850 mg/kg bw/d | Oral perfusion | GD14-18 | PND1, and 18 months of age | Renal development and fibrosis | Bodyweight, kidney size, and kidney-to-bodyweight ratio: DBP↓ (PND1) | Swelling of the glomerular tufts: DBP ↑ (18 months) | [261] |
| Rat | DBP | 850 mg/kg bw/d | Oral perfusion | GD12-18 | PND70 | Renal development and fibrosis | Bodyweight: DBP ↓ (♂) | Kidney Fgf10, Fgfr2, and AR protein levels: DBP ↓ (♂) | [262] |
| Rat | DBP | 850 mg/kg bw/d | Oral gavage | GD14-18 | PND1 and 8 weeks of age | Renal development and fibrosis | Fetal bodyweight and survival rate: DBP ↓ | Mother and fetal morbidity: no effect | [263] |
| Rat | DEHP, DBP | 7, 70, or 700 mg/kg bw/d DEHP—5, 50, or 500 mg/kg bw/d | Oral gavage | GD13—PND21 | PND4, 21, 22, 33, 78, 88, and 90 | Metabolism | Bodyweight from PND16 to 21: DEHP ↑ (700)/DBP ↑ (5) | Dams bodyweight gain and organ weight: no effect | [243] |
| Mouse | DEHP, DiNP | DEHP 25 mg/kg chow/d, DiNP 75 mg/kg chow/d or both | Oral route (food) | GD-15 to PND21 | PND21 and 10 months | Liver metabolism | Number of DEGs at PND21: DiNP (♀ (61))/DEHP + DiNP (♀ (1)) | [237] | |
| Rat | n-BuP | 10, 100, or 500 mg/kg bw/d | Oral gavage | GD7-PND22 | PND1, 6, 14, 16, 17, 22, and 80-90 | Fetal development/Reproduction | Fetal AGD: n-BuP ↓ (10, 500) | Maternal bodyweight and gestational length: no effect | [264] |
| Rat | n-BuP | 850 mg/kg bw/d | Intragastric administration | GD14-18 | PND1 | Renal development | Kidney autophagy marker (LC3B, Beclin-1) staining and expression: n-BuP ↑ | [265] | |
| Mouse | n-BuP | 3.5 µg/week/mouse | Subcutaneous injection | GD0-PND28 | PND28–PND84 | Adipogenesis and food intake regulation | Bodyweight, fat mass, weekly food intake, fasting serum glucose, and leptin level: n-BuP ↑ (♀) | 17β estradiol levels: no effect | [234] |
Results in brackets correspond to the sex in which the effect was observable (♂, male; ♀, female), followed by stage at observation separated by a column from the doses administered at which the effect was observed, and, if so, followed by the number of DEGs in brackets. If one of the above-mentioned characteristics is not specified, the effect was observed for both sexes, at the different stages of observation and at the different doses used. Arrows indicate the direction of the effect: a downward arrow (↓) indicates a decrease and an upward arrow (↑) indicates an increase. ABCG5: ATP-binding cassette transporter G5; ACC1: acetyl-CoA carboxylase 1; Acly: ATP-citrate lyase; ACSL1: acyl-CoA synthetase long-chain family member 1; Adh1: alcohol dehydrogenase 1; Adipoq: adiponectin; Adrb3: adrenoceptor beta 3; AGPAT1: 1-acylglycerol-3-phosphate O-acyltransferase 1; Agrp: agouti-related protein; AGTR1A: angiotensin II receptor type 1A; AGTR1B: angiotensin II receptor type 1B; AGTR2: angiotensin II receptor type 2; Agtrap: angiotensin II receptor-associated protein; ALAT: alanine aminotransferase; ALB: albumin; ANP: atrial natriuretic peptide; Aox1: aldehyde oxidase 1; AR: androgen receptor; ASAT: aspartate aminotransferase; AT1R: angiotensin II receptor type 1; AT2R: angiotensin II receptor type 2; Atf4: activating transcription factor 4; Atf6: activating transcription factor 6; Bax: BCL2-associated X apoptosis regulator; Bcl-2: BCL2 apoptosis regulator; Bcrp: breast cancer resistance protein; Bip: heat shock protein family A (Hsp70) member 5 (or Hspa5); Bmp4: bone morphogenetic protein 4; n-BuP: butylparaben; BPA: bisphenol A; BPF: bisphenol F; BPS: bisphenol S; BUN: blood urea nitrogen; bw: bodyweight; Calm1: calmodulin 1; CCDC152: coiled-coil domain containing 152; CCL2: chemokine (C-C) ligand 2; Cdh11: cadherin 11; Cidea: cell death-inducing DFFA-like effector a; COL1A1: collagen-1a1; COLEC12: collectin subfamily member 12; CPT1α: carnitine palmitoyl transferase 1 alpha; CRTC2: CREB-regulated transcription coactivator 2; Cs: citrate synthase; Cyp1a2: cytochrome P450 family 1 subfamily A member 2; CYP7A1: cholesterol 7α-hydrolase; Cyp11b2: cytochrome P450 family 11 subfamily b: polypeptide 2; Cyp17: steroid 17-alpha-hydroxylase; Cyp19: cytochrome P450 family 19; Cyp19A1: cytochrome P450: family 19: subfamily a: polypeptide 1; d: day; D1: cyclin D1; DBP: dibutyl phthalate; DEHP: di(2-ethylhexyl) phthalate; DGAT1: diacylglycerol O-acyltransferase 1; DiNP: di-isononyl-phthalate; DNA: deoxyribonucleic acid; DNMT1: DNA methyltransferase 1; DNMT3A: DNA methyltransferase 3 alpha; DNMT3b: DNA methyltransferase 3 beta; ECE1: endothelin converting enzyme 1; EFAs: essential fatty acids; eNOS: endothelial nitric oxide synthase; Esr1: estrogen receptor 1; ESRRA: estrogen-related receptor alpha; F1: first generation: F2: second generation; F3: third generation; FA: fatty acid; Fabp4: fatty acid-binding protein 4; Fads: fatty acid desaturase 1; FAS: fatty acid synthase; Fgf10: fibroblast growth factor 10; Fgf21: fibroblast growth factor 21; Fgfr2: fibroblast growth factor 2; Foxd1: forkhead box D1; FoxO1: forkhead box O1; FSH: follicle-stimulating hormone; G-6-Pase: glucose-6-phosphatase; GD: gestational day; Gdnf: glial cell-derived neurotrophic factor; GJA5: gap junction protein alpha 5; Gli1: GLI family zinc finger 1; GLUT2: glucose transporter 2; GLUT4: glucose transporter 4; Gnpat: glyceronephosphate O-acyltransferase; GPAT: glycerol-3-phosphate acyltransferase; Gpd1: glycerol-3-phosphate dehydrogenase 1; GSEA: gene set enrichment analysis; GSH-px: glutathione peroxidase; GSK3β: glycogen synthase kinase 3 beta; GSR: glutathione reductase; HBB: hemoglobin subunit beta; HDLs: high-density lipoproteins; HhIP: hedgehog-interacting protein; Hnf1a: hepatocyte nuclear factor 1 homeobox A; HNF4-α: hepatocyte nuclear factor 4 alpha; IFN-γ: interferon gamma; IL-1β: interleukin 1 beta; IL-2: interleukin 2; IL-4: interleukin 4; IL-5: interleukin 5; IL-6: interleukin 6; IL-12p70: interleukin 12 heterodimer of p40 and p35 subunits; IL-13: interleukin 13; IL-17; interleukin 17; IL-21: interleukin 21; IL-23: interleukin 23; IsoBP: isobutylparaben; LC3B: microtubule-associated protein 1 light-chain 3 beta; LC-PUFAs: longer-chain polyunsaturated fatty acids; LDLs: low-density lipoproteins; LDLR: low-density lipoprotein receptor; lepr: leptin receptor; LH: luteinizing hormone; lncRNA: long noncoding ribonucleic acid; Lpl: lipoprotein lipase; LPS: lipopolysaccharide; LRP-1: LDLR-related protein 1; MafA: MAF bZIP transcription factor A; mc4r: melanocortin type 4 receptor; MDB2: methyl-CpG binding domain protein 2; MeCP2: methyl CpG binding protein 2; MGAT1: alpha-1,3-mannosyl-glycoprotein 2-beta-N-acetylglucosaminyltransferase 1; miRNA: microribonucleic acid; MMP7: matrix metallopeptidase 7; mRNA: messenger ribonucleic acid; Mrp3: multidrug resistance-associated protein 3; MSLN: mesothelin; mTOR: mammalian target of rapamycin; MUFAs: monounsaturated fatty acids; NIS: sodium iodide symporter; NP: 4-nonylphenol; nPE1: POMC neuronal enhancer; Npy: neuropeptide Y; OP: 4-tert octylphenol; Pax2: paired box 2; Pax4: paired box 4; Pax6: paired box 6; PAX8: paired box 8; Pck1: phosphoenolpyruvate carboxykinase 1; Pdx1: pancreatic and duodenal homeobox 1; PEPCK: phosphoenolpyruvate carboxykinase; PLS-DA: partial least-squares discriminant analysis; PND: postnatal day; PNPLA2: patatin-like phospholipase domain-containing 2; Pparg: peroxisome proliferator-activated receptor gamma; Ppargc1b: PPARγ coactivator 1; PRKCE: protein kinase C epsilon; Ptch1: patched 1; PUFAs: polyunsaturated fatty acids; RhoA: Ras homolog gene family member A; ROCK1: Rho-associated protein kinase 1; RORγt: retinoic acid-related orphan receptor t; Rxr-β: retinoid X receptor β; Rxrg: retinoid X receptor gamma; pomc: pro-opiomelanocortin; SCD-1: stearoyl-CoA desaturase 1; Slc2a2: solute carrier family 2 member 2; Slc19a2: solute carrier family 19 member 2; SLC44A4: solute carrier family 44 member 4; SMAD7: suppressor of mothers against decapentaplegic 7; SPRY1: sprouty RTK signaling antagonist 1; SR-BI: scavenger receptor class BI; Srebf1: sterol regulatory element-binding transcription factor 1; SREBP-1: sterol regulatory element-binding protein-1; snoRNA: small nucleolar ribonucleic acid; TC: total cholesterol; Tcf21: transcription factor 21; TCS: triclosan; TGs: triglycerides; TIF-1: thyroid transcription factor 1; TP: total protein; Ucp1: uncoupling protein 1; UDPGT: UDP glucuronosyltransferase; WFDC2: WAP four-disulfide core domain 2; Wnt11: Wnt family member 11; Ywhab: tyrosine 3-monooxygenase/tryptophan 5-monooxygenase activation protein beta.
6.2. Effects of PPPs on Cardiometabolic Disorders
6.2.1. Epidemiological Data
Most studies examining environmental pollutants have reported associations with increased blood pressure in children. However, the literature reported only putative evidence for the effects of exposure to phenols or phthalates [239]. In human cohorts, the cardiovascular characteristics of the offspring after maternal exposure to PPPs have been poorly examined due to difficulties in assessing blood pressure in infants. Nevertheless, blood pressure was measured in 1277 children aged 6 to 11 years from the European Human Early-Life Exposome (HELIX) cohort. Increases in diastolic blood pressure in infants were observed with prenatal BPA exposure. Interestingly, blood pressure was also measured in 1015 Spanish preadolescents (mean age = 10.4 years), in whom prenatal exposure to BP-3 was associated with higher diastolic blood pressure [174].
Regarding parabens, a decrease in systolic blood pressure in preadolescents was observed with MeP, mainly in boys, and with n-BuP only in girls [174].
Regarding phthalates, prenatal exposure to these compounds was associated with lower systolic blood pressure at ages 4 and 7 in girls, but not boys [191], and with lower cholesterol levels at 8–14 years for both [266]. Vafeiadi et al. also observed a negative association between phthalate metabolite concentrations and systolic/diastolic blood pressure at 4 years in both sexes [198]. Among the few studies examining the association between prenatal phthalate exposure and children’s blood pressure, two studies reported a negative association between ∑DEHP and systolic/diastolic blood pressure [198] in the previously described European HELIX cohort, and between systolic blood pressure and MBzP [267]. Additionally, high- and low-molecular-weight phthalate mixtures were associated with a decreased metabolic score, which included waist circumference, systolic and diastolic blood pressure, triacylglycerol, high-density lipoprotein cholesterol (HDL), and insulin levels [202].
Most studies have highlighted the effects of phenols and phthalates on maternal health, child growth, and cardiometabolic outcomes, but the results are conflicting. Cardiometabolic risks in adulthood need to be confirmed and supported by further rigorous studies [268].
6.2.2. Maternal Exposure to PPPs and Cardiovascular Outcomes in Experimental Models
A growing body of evidence suggests that early life exposure to BPA may also have a substantial influence on perinatal and postnatal cardiometabolic programming, contributing to higher cardiometabolic risks in adulthood [249].
In sheep, prenatal BPA exposure (0.5 mg/kg BW/day by injection for the last two-thirds of gestation) had no significant effect on blood pressure or cardiac morphometric measurements in the offspring. However, this exposure increased the atrial natriuretic peptide gene expression in the ventricles and reduced the collagen expression in the right ventricle. When the mothers are overfed, BPA amplifies septal hypertrophy and continues to block left ventricular hypertrophy and blood pressure. Prenatal BPA seems to thwart obesity-induced cardiovascular disorders [251]. The exposure of female Sprague–Dawley rats during the last two-thirds of gestation to a mixture of BPA (0.005 mg/kg bw/day) and a high dose of DEHP (7.5 mg/kg bw/day) produced postnatal outcomes, including increased relative heart weight in adult male offspring [158]. In mice, lifelong maternal exposure to DEHP (30 mg/kg bw through daily oral administration) until weaning had the following effects on 8-week-old offspring: (i) increased blood pressure, (ii) deregulated aortic eNOS (endothelial nitric oxide synthase) phosphorylation, and (iii) upregulated AT1R (angiotensin I receptor) protein expression (angiotensin II signaling) [236].
Additionally, this perinatal exposure to DEHP was shown to increase adiposity in the offspring with increased bodyweight and WAT- and BAT (white and brown adipose tissue)-to-body-weight ratios, along with impaired hepatic cholesterol metabolism (increased plasma and hepatic cholesterol). Linked to the cardiovascular effects of perinatal exposure to PPPs in offspring, a study in mice highlighted the potential role of DNA methylation in DEHP-induced cardiac effects and emphasized the importance of gender/sex as a biological variable in environmental health studies [253]. Exposure to DEHP (300 mg/kg bw/day) during the last third of gestation, reduced systolic and diastolic systemic arterial blood pressures, and locomotor activity, in PND200 male rats. This exposure to DEHP decreased aldosterone release in these males [259], while aldosterone increased in females. This suggests a sex-specific adrenal response to in utero exposure to DEHP and opens up the possibility of a hypertensive response induced by DEHP in the female offspring [256,257]. However, maternal exposure to a lower dose of DEHP (10 mg/kg bw/d) was shown to induce hypertension and bodyweight gain in male offspring [269]. Another study investigating the effects of maternal DEHP exposure during whole gestation showed impaired renal development in offspring. This led to a nephron deficit, and subsequently elevated blood pressure later in life, by the inhibition of the renin–angiotensin system [260]. Maternal exposure to DBP (850 mg/kg bw/day during GD14-18) induced kidney dysplasia and renal fibrosis in male offspring [261] by an expression of TGF-β1 and the abnormal activation of the epithelial–mesenchymal transition in fibrotic kidneys [270]. This exposure also decreased the testosterone concentration and reduced androgen receptor expression [262]. Maternal exposure to a lower dose of DEHP (10–100 mg/kg bw/d) induced a decreased heart weight and altered cardiac metabolic function in young offspring [258]. Maternal exposure to phthalic acid reduced the bodyweight (bw), heart weight (HW), and HW/bw in offspring, while their heart rates and blood pressures were conversely increased compared to the control group [271]. In the Sprague–Dawley rats, maternal exposure to DBP (850 mg/kg/day orally in the last third of gestation) induced renal fibrosis in the offspring [263,265].
These data highlight the difficulties in establishing a link between PPP exposure (alone or in combination or mixture) and the risk of metabolic and cardiovascular diseases. This is why it is necessary to develop new approaches using a new animal model to assess the long-term effects of exposure to PPPs, including at puberty and in adulthood, and taking into account the sex of the offspring.
6.3. Effects of PPPs on Gonadal Functions and Fertility in Cohorts and Animal Models
The gonads are the primary organs of the reproductive system responsible for the production of sex hormones and gametes. Thus, the gonads are particularly susceptible to endocrine disruptors, which are known to interfere with steroid hormone receptors. These include the estrogen receptor (ER), progesterone receptor (PR), and androgen receptor (AR), causing reproductive dysfunctions. Therefore, this part of the paper will be dedicated to the effect of maternal exposure to BPA, TCS, parabens, and phthalates or a mixture of pollutants on the gonadal functions of male and female offspring in cohorts and/or animal models.
6.3.1. Epidemiological Data
Although difficult to undertake due to the duration of follow-up, the effects of prenatal exposure to PPPs on human reproduction are essential to understand the short-term and long-term effects on offspring [272].
One of the recognized markers of prenatal androgenization widely used and easily accessible is the measurement of anogenital distance [273]. In newborn boys (n = 72), but not in girls, a significant association between high cord blood BPA levels and shortened anoscrotal distance was suggested by Mammadov [274]. Similarly, boys whose mothers had detectable levels of BPA in their urine at 12–16 weeks of gestation were more likely to have shorter anogenital distance at birth, and at both at 6 and 12 months of age, but not in girls [275].
Related to puberty, the CHAMACOS cohort study reported an association of higher prenatal exposure to BPA with later puberty in girls and earlier puberty in boys [276]. In contrast, maternal urinary concentrations of BPA from three birth cohorts, the INMA (Spain), EDEN (France), and MoBa (Norway), were associated with delayed pubertal development in boys and girls [277]. In the study by Ferguson et al., prenatal exposure to BPA was associated with decreased odds of adrenarche (maturation of adrenal androgen secretion) and puberty in boys between 8 and 14 years of age [278].
To our knowledge, only two studies have investigated the association between the in utero exposure to phenol and reproduction in men in adulthood [279]. In the Western Australian Pregnancy Cohort, maternal BPA exposure at 18 and 34 weeks of gestation was positively associated with their sons’ sperm concentration and motility in adulthood (20–22 years), but not for their testicular function [279]. Furthermore, a high exposure to BPA and BP-3 during gestation was characterized by a compensated reduced Leydig cell function in adulthood, with no association with anogenital distance or semen quality [280].
Concerning prenatal TCS exposure, one study has explored its impact on reproductive hormones in cord blood, where TCS was associated with an increase in T and a decrease in E2 concentrations in cord blood among male infants [281]. Earlier puberty has been suggested in girls in all three birth cohorts, the INMA (Spain), EDEN (France), and MoBa (Norway) [277].
Later on, earlier menarche was reported with prenatal TCS exposure in girls from the CHAMACOS cohort, focusing on Latino children [282]. Using the same cohort, earlier breast development, pubic hair development, and menarche were observed with prenatal methylparaben (MeP) exposure in girls, and only earlier menarche with propylparaben. In boys, only genital development was reported associated with prenatal propylparaben exposure [282]. In addition, n-BuP was negatively associated with pubertal onset, adrenarche, and/or gonadarche (earlier gonadal development) in boys in the three European cohorts, the INMA (Spain), EDEN (France), and MoBa (Norway). In contrast, EtP and PrP were negatively associated with gonadarche in boys [277].
The association between prenatal phthalate exposure and gonadal development has been studied. Swan et al. reported an association between the anogenital distance, the incomplete testicular descent, the incomplete virilization, and the prenatal phthalate monoester metabolites exposure [283]. In the TIDES cohort from the USA, first-trimester urinary DEHP metabolite concentrations were associated with increased odds of genital abnormalities, especially due to hydrocele in newborn males [284]. In contrast, in a Canadian pregnancy cohort study, no strong evidence was observed between maternal urinary phthalates in the first trimester and the length or width of the penis at birth [285]. The association of prenatal exposure to phthalates with pubertal timing in boys and girls was investigated using several cohorts. Prenatal exposure to most phthalates was associated with reduced odds of adrenarche (notably with MEHHP, MEOHP, and MBzP) and reduced odds of puberty in association with DEHP metabolites, MBzP, MBP, and MiBP [278].
In the CHAMACOS longitudinal cohort study, it was found that the cx-MiNP, MCOP, and MCPP were associated with the late onset of pubarche (the appearance of pubic hair) and menarche (the first period) mostly among normal-weight girls, whereas MBzP was associated with later thelarche in girls. In boys, all of the phthalate biomarkers were associated with earlier gonadarche and pubarche, but associations with phthalate metabolites were close to the null or positive [276]. Moreover, prenatal urinary MEP concentrations were associated with an earlier onset of pubic hair development only in girls [282].
In the INMA Spanish cohort study, prenatal exposure to DEHP was associated with a higher risk of puberty onset and gonadarche in boys, and a higher risk of adrenarche in girls, at age 7–10 years. In boys, prenatal exposure to DEP, DnBP, and DEHP was also associated with a higher risk of adrenarche or gonadarche in children with normal weight, and BBzP and DiNCH exposure with lower risk of adrenarche in children with overweight/obesity. In girls, DiBP, DnBP, and DiNCH were associated with a higher risk of gonadarche in children with overweight/obesity [286]. A recent publication from Freire et al. reported that MEHP was associated with delayed gonadarche and adrenarche, respectively, in girls and boys [277].
Interestingly, associations between prenatal exposure to phthalates and peripubertal measures of male reproductive development seem to be dependent on the timing of in utero exposure. In fact, exposure to phthalates during the third trimester was associated with reduced odds of having a Tanner stage higher than one for pubic hair development and higher peripubertal SHBG (sex hormone-binding globulin) levels. Exposure to phthalates during the first- and second-trimester phthalates was not associated with these outcomes. In contrast, only during the first trimester, exposure to DEHP was associated with higher estradiol concentration [287]. The long-term effects on adult reproductive health in male offspring (18–20 years) were also investigated, demonstrating an association between high maternal exposure to some phthalates and an impaired Leydig cell function characterized by lower total and free testosterone/LH ratios [288].
All of these studies suggested that PPPs can affect male and female genital development, puberty, semen quality, and sexual hormones, but the effects can depend on the window of prenatal exposure.
6.3.2. Maternal Exposure to PPPs and Gonadal Effects Using Animal Models
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Male Offspring
Most studies that have assessed the risk of perinatal exposure to BPA on male reproductive function have shown abnormal production of sex hormones in adulthood. In rats, daily exposure to 0.025 mg/kg of bodyweight or 0.250 mg/kg bw of BPA during the second half of gestation (GD10-21) reduced circulating testosterone levels (T) and luteinizing hormone (LH). This exposure also increased the levels of follicle-stimulating hormone (FSH) and estradiol (E2) [289]. Similar effects on blood levels of T and E2 are observed when exposure begins at the start of gestation (0.010 mg/kg bw) [290]. However, the opposite effects were observed for these hormones when measured at weaning [290]. Similarly, perinatal exposure (from GD18 to PND5) to 0.5 or 5 mg/kg bw of BPA results in a higher production of T [291]. In addition, exposure to BPA induces a decrease in AR expression and an increase in ER expression [291,292]. As a result, multiple abnormal histological and architectural damages are seen during testicular development. These include interstitial necrosis, germ cell degeneration, decreased tubular and luminal diameter [289,291,293], acrosome and plasma membrane integrity alteration, decreased mitochondrial activity [291], and increased oxidative stress [289]. Additionally, daily exposure to low levels of BPA (in the range of 0.0012 mg/kg bw and 5 mg/kg bw) significantly reduces sperm motility and count [289,291,292]. These outcomes lead to alterations in the fertility of the first male offspring and their subsequent second and third generations. However, this exposure does not affect the transcriptomic and epigenomic profile at the genome scale in sperm [292,293].
Studies conducted using mice or rabbits as a model have revealed similar morphological and hormonal damages in the testis of offspring exposed to BPA [248,294,295]. The observed effects on steroidogenesis are associated with a dysregulation of the steroidogenic enzyme expression. For example, daily exposure to BPA at 500 mg/kg bw from GD8 to PDN14 in mice decreased the expression of StAR (Steroidogenic Acute Regulatory Protein) and Cyp11a (Cytochrome P450 Family 11 Subfamily A Member 1) [296]. In rabbits, daily exposure to BPA at 50 mg/kg bw from GD15 to birth decreased the expression of CYP11A1 and 3β-HSD [297]. However, these effects are associated with different effects on T levels, which are reduced in adult mice and increased in 3-day-old rabbits. In addition, other transcriptomic studies showed that the expression of Snrnp 40 (U5 small nuclear ribonucleoprotein subunit) was upregulated in the spliceosome pathway and that the expression of Hnrnpu (encoding a DNA- and RNA-binding protein) was downregulated. These expressions suggest that spliceosome blockage may be the cause of abnormal testicular development in male mice exposed to BPA [295]. On the other hand, prenatal BPA exposure upregulated the transcription level at PND21 of testicular Dnmt1 and inhibited the transcription of testicular Dnmt3A and Dnmt3B. In addition, the transcriptional level of testicular caspase-7, caspase-9, and bax is increased, and the transcriptional level of bcl-2 is decreased at PND56, leading to apoptosis in the testis [295].
Gestational and postnatal exposure to phthalates also induced adverse effects in male offspring. In rabbits, daily exposure to DBP (400 mg/kg bw) in utero (GD15-29) reduced serum T levels and, consequently, the concentration and quality of ejaculated sperm and the weights of the testes and accessory sex glands. Interestingly, rabbits exposed to DBP developed cryptorchid testes with carcinoma in situ-like cells, malformed foreskin (giving the appearance of feminized external genitalia), hypospadias, hypoplasia and prostate, and agenesis of the bulbourethral gland [298]. In addition, gestational daily exposure (GD10-20) to DEHP equivalent to a human daily intake (0.0024–0.003 mg/kg of bw/day) in mice reduced the anogenital distance, seminal vesicle weight, expression of testicular steroidogenic enzymes (Star, Cyp17a1, Hsd17b12, Hsd3b1, and Hsd3b6), and sperm count. However, this exposure did not affect either the testicular morphology or fertility performances [299]. It was previously assumed that the effect of DEHP on T production and steroidogenic enzyme expression was strain-dependent in rats [300]. However, Hannas and al. reported very minor differences in the testicular and epidydimal phenotype between the Sprague–Dawley and Wistar rat strains [301].
Over the years, the use of animal models to understand the effects of maternal exposure to PPPs during pregnancy on the offspring phenotype has provided evidence that this exposure caused genital abnormalities in both male and female offspring. These studies were conducted by exposing animals to single compounds. However, on a daily basis, humans are exposed to dozens or even hundreds of chemical combinations through inhalation, skin contact, and ingestion. Consequently, the number of studies using mixtures of toxic substances is increasing. For example, an epidemiological study including 194 pregnant women (European EDC-MixRisk project (
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Female Offspring
As for the male offspring, the harmful effects on female reproductive functions are induced by gestational or perinatal exposure to BPA, phthalates, or a mixture of pollutants. However, a limited number of studies focusing on female reproductive health have been published in the past five years.
In mice, gestational exposure to BPA (2.5, 5, 10, 20, and 40 mg/kg bw/day from GD0.5 to GD17.5) advanced puberty, induced atrophy of the ovary at adulthood, and female offspring were more likely to abort [307]. At higher doses, BPA (50, 500, and 2500 mg/kg bw) caused ovarian damage by increasing vacuole formation and decreasing the number of corpus granules [290,296]. In rats, prolonged exposure to BPA through drinking water (1 μg/mL or 10 μg/mL BPA from GD6 to PND21) caused similar damage to female sexual differentiation, such as advanced puberty and endometrial malformation [296,308]. Multiple hypotheses of the BPA action mechanism have been suggested. BPA has been shown to exert its effect by binding to the ER and mimicking the weak effect of estrogen. Interestingly, serum levels of E2, and ovarian ERa and ERb, were reduced in female offspring exposed in utero upon reaching adulthood. On the other hand, a decrease in the ovarian expression of Dnmt1, Dnmt3A, Dnmt3B, and Bcl2, and an increase in the relative expression of caspase-9, caspase-7, and bax, were seen in female offspring exposed to BPA [290,296,307]. Moreover, the activation of inflammation and abnormal autophagy via the TLR4/NF-κB and mTOR signaling pathways have been reported in the ovary and in the uterine tissue of exposed female offspring [308]. Altogether, these observations suggest that the exposure to BPA leads to the dysregulation of estrogen levels, altered DNA methylation, and the activation of inflammation and apoptosis in the ovaries of females of the next generation.
In females, the effects of maternal exposure to DEHP have negative effects on oocyte growth, meiotic maturation, and ovarian function [309]. Specifically, in first-generation ovarian primordial germ cells (GD12.5) and first- and second-generation oocytes (PND21), maternal exposure to DEHP (0.040 mg/kg bw/day, from GD0 to birth) decreases the methylation of CpG sites in the maternal imprinted gene Igf2r (insulin-like growth factor 2 receptor) and in the paternal imprinted gene Peg3 (paternally expressed gene 3) [310]. Additionally, Pocar et al. showed that a longer exposure to a higher level of DEHP (0.05 or 5 mg/kg bw/day, from GD0 to PND21) decreased the expression of Cyp19a1 and Cyp17a1, and blocked the process of meiosis II in most oocytes after superovulation in adult female offspring [311]. More recently, Repouskou et al. reported a cumulative effect of a mixture of phthalates. Pregnant mice exposed to a mixture of four phthalate monoesters (33% MBP, 16% MBzP, 21% MEHP, and 30% MiNP at 0, 0.26, 2.6, and 13 mg/kg bw/day from GD0 to birth) reduced the number of preantral follicles (primary and secondary), increased follicular atresia, and reduced the Cyp19a1 expression in the ovaries of female offspring [302].
Altogether these studies conducted in animal models demonstrated that PPPs can affect gonadal development, puberty, steroidogenic enzymes, sexual hormones, semen, and oocyte quality.
7. Conclusions
Despite the implementation of regulations aimed at banning certain PPPs or reducing their concentrations in consumer products, the general population, including pregnant women, continues to be exposed daily to these pollutants. The objective of this review was to compile the current knowledge on the effects of prenatal exposure to these molecules on fetoplacental development and offspring health in the context of the developmental origins of health and disease concept (DOHaD).
Due to a short half-life, these molecules are rapidly eliminated in the urine in the form of one or more metabolites. Maternal urine sampling appears relevant for biomonitoring the exposure to these molecules. Worth noting, while spot urine samples are of interest in large populations for describing exposure, their use in etiologic studies is limited. Given the reported high intra-individual variability in urinary concentrations of some phenols and phthalates, these snapshot assessments imperfectly reflect the average exposure over a long period, such as pregnancy. This leads to measurement errors and estimated effects biased toward zero. As shown in a simulation study, for a compound with a high intra-individual variability, such as BPA (intraclass correlation coefficients of around 0.2), the bias in effect estimates can be as high as 80%. The collection of repeated urine samples in the period of interest is, therefore, necessary to properly assess the exposure to PPPs and their effects on human health [312]. Unfortunately, this approach is not always applied to cohorts due to logistical constraints. Additionally, preconception maternal exposure levels are rarely known, as expectant mothers are recruited into cohorts once their pregnancies have been established.
During gestation, PPPs from the maternal bloodstream are able to cross the placental barrier in a native form or as metabolites and reach the fetal circulation, thereby contributing to direct exposure to the fetus. In humans, depending on the compound, positive or negative associations have been observed with placental weight or newborn weight and size, with, in some cases, sex-specific effects. Some studies have also found associations with DNA methylation, but generally without exploring the link between these methylation patterns and gene expression or any biological significance.
In vitro and in vivo studies were used to investigate the individual effects of PPPs on placental function. In vitro, depending on their concentration, according to a non-monotonic dose–response relationship, i.e., a U-shaped curve, phenols may or may not affect the proliferation of placental trophoblast cells and their hormone production. On the other hand, with monotonic dose–response curves, i.e., progressive effects depending on their concentrations, phthalates show effects on the lipid content, hCG secretion, and cell fusion through PPARγ. The in vitro approach makes it possible to test several concentrations of PPPs, alone or in a mixture, and to decipher their mechanism of action, but it does not take into account the sub-chronic effects or the complexity of the different cell types that make up the placenta. It is, therefore, essential to complement these studies with in vivo models.
The use of different animal models made it possible to evaluate the effects of sub-chronic exposure, to test several concentrations, and to target different periods of gestation. BPA appears to affect the placental structure, the expression of hormone receptors, and the genes involved in DNA methylation, as well as the DNA methylation levels, while TCS affects the expression of nutrient transporters and hormones. Phthalate-related alterations have been reported in placental morphology, hormone production, vascularization, histopathology, and gene/protein expression [131]. It should be noted, however, that the effects of pollutants on placental function are never evaluated in a cocktail, and that the exposure time only partially covers the preconception and/or gestation periods. Furthermore, sex-specific effects are rarely taken into account in these studies.
In a DOHaD context, it is also important to take into account the effects of prenatal exposure on the postnatal phenotype to assess population risks and, if necessary, adapt the PPP regulations. In human cohorts, depending on the children’s sex, positive or negative associations have been established between prenatal exposure to PPPs and children’s BMI, blood pressure, gonadal function, or age at puberty. However, the phenotyping of offspring is sometimes limited in terms of exploratory physiological tests for obvious ethical reasons, and, given the intergenerational duration, postnatal monitoring is currently limited to puberty. It is, therefore, essential to establish the postnatal phenotype of the offspring reaching adulthood, and even transgenerationally, by generating animal models exposed to PPPs at different doses and exposure durations so to complement the data from human cohorts.
Depending on the level of in utero exposure, and on the age and sex of the offspring, bisphenols and phthalates have been shown to affect bodyweight, carbohydrate homeostasis, insulin sensitivity, and thyroid hormones in offspring. These pollutants also have cardiovascular effects. Bisphenols can lead to ventricular hypertrophy, while phthalates affect blood pressure, with the phenotypes depending on the sex. As PPPs are endocrine disruptors, these molecules disrupt gonad development, the age of puberty, sex hormone concentrations, and the quality of sperm and oocytes.
Future challenges will be to define a pollutant mix based on data from human cohorts, including fetoplacental biometric data and postnatal phenotyping data representative of the maternal exposome. The next challenge will be to study the effects of maternal exposure by the ingestion of this mixture of PPPs in an animal model mimicking human maternal exposure. In particular, this challenge needs to explore the effects on the development and fetoplacental growth, as well as on the postnatal phenotype, especially the cardiometabolic status and gonadal function, whilst taking into account the sex of the offspring. The question of the animal model is crucial. It is important to choose a model whose placenta is close to that of the human placenta.
Currently, rodent models are the most widely used. However, their placenta is made up of three layers of trophoblastic cells and is therefore more distant from the human placenta than that of lagomorphs [313]. This makes lagomorphs an interesting model to explore. Additionally, the establishment of placental epigenetic signatures seems to be a relevant issue to link to the postnatal phenotype, and thus to predict the health trajectory of the offspring. These data could then be extrapolated to follow-up cohorts to assess the potential phenotypic risks for children exposed in utero to PPPs. By doing so, preventive measures can be put in place to limit the exposure to PPPs, and regulations could be revised, if necessary.
Conceptualization, D.R.-R., V.G., C.P. and A.C.-T.; writing—original draft, D.R.-R., J.B., M.O., N.J., V.G., N.M., M.-N.D., N.P.-H., M.-J.F.-S., H.J., C.P. and A.C.-T. Writing—review and editing, A.C.-T. and D.R.-R., Funding acquisition, A.C.-T. All authors have read, improved, and approved the final version of the manuscript. All authors have read and agreed to the published version of the manuscript.
Not applicable.
Not applicable.
Not applicable.
The authors declare no conflicts of interest.
AR: androgen receptor; AT1R: angiotensin 1 receptor; ATM: ataxia telangiectasia mutated; BAT: brown adipose tissue; BBP: benzyl butyl phthalate; Bcl-2: B cell lymphoma 2; 11beta-HSD2: 11-beta-hydroxysteroid dehydrogenase 2; BMI: body mass index; BMPP: bis (4-methyl-2-pentyl) phthalate; n-BuP: n-butylparaben; BP-1: benzophenone-1; BP-3: benzophenone-3; BP-8: benzophenone-8; BPA: bisphenol A; BPAF: bisphenol AF; BPAP: bisphenol AP; BPB: bisphenol B; BPC: bisphenol C; BPE: bisphenol E; BPF: bisphenol F; BPFL: bisphenol FL; BPM: bisphenol M; BPP: bisphenol P; BPS: bisphenol S; BPZ: bisphenol Z; bw: bodyweight; BzP: benzylparaben; CCL2: C-C motif chemokine ligand 2; CL: blood clearance; CRE: CBP-responsive element; CREB: cAMP-responsive element-binding protein; CRH: corticotrophin-releasing hormone; CSS: steady-state plasma concentration; CTR1: SLC31A1 for solute carrier family 31, member 1; cx-MiNP: mono(carboxy-iso-nonyl) phthalate; DBEP: bis (2-n-butoxyethyl) phthalate; DBP: di-n-butyl phthalate; DCHP: dicyclohexyl phthalate; 2,4-DCP: 2,4-dichlorophenol; 2,5-DCP: 5-dichlorophenol; DEEP: bis (2-ethoxyethyl) phthalate; DEHP: di-(2-ethylhexyl) phthalate; DEP: diethyl phthalate; DnHP: di-n-hexyl phthalate; DI: daily PPP intake; DiBP: diisobutyl phthalate; DiDP: di-isodecyl phthalate; DiNCH: di-iso-nonyl-cyclohexane-1,2-dicarboxylate; DiNP: di-isononyl phthalate; DMEP: bis(2-methoxyethyl) phthalate; DMP: di-methyl phthalate; DnBP: di-n-butyl phthalate; DNMT1: DNA (cytosine-5)-methyltransferase 1; DnOP: di-n-octyl phthalate; DNP: dinonyl phthalate; DPP: di-amyl phthalate; E2: estrogen; ECHA: European Chemicals Agency; EDCs: endocrine-disrupting chemicals; EFSA: European Food Safety Authority; EGF: epidermal growth factor; eNOS: endothelial nitric oxide synthase; EtP: ethylparaben; ER: estrogen receptor; ERE: estrogen response element; F: bioavailability; FPR: birthweight-to-placental-weight ratio; FSH: follicle-stimulating hormone; Fue: urinary excretion factors; GCNF: Germ cell nuclear factor; GLUT-1: glucose transporter 1; GWs: gestational weeks; hCG: human chorionic gonadotropin; HEPH: hephestin; HIF-1α: hypoxia-inducible factor 1-α; HMW: high-molecular-weight phthalates; HeP: heptylparaben; Hsp70: heat shock protein 70; 5-HT: serotonin; ICR: imprinting control region; IGF-1 or -2: insulin-like growth factor; IGF-2R: insulin-like growth factor-2 receptor; IGFBPs: IGF-binding proteins; IL: interleukin; i-BuPB: iso-butylparaben; i-PrPB: iso-propylparaben; LH: luteinizing hormone; LMW: low-molecular-weight phthalates; LXR: liver X receptor; MBP: mono-n-butyl phthalate; MBzP: monobenzyl phthalate; MCIOP: mono(4-methyl-7-carboxyheptyl) phthalate; MCMHP: mono-2-carboxy-methyl hexyl phthalate; MCOP: mono-carboxy-iso-octyl phthalate; MCPP: mono(3-carboxypropyl) phthalate; MECPP: mono(2-ethyl-5-carboxy-pentyl) phthalate; MECPTP: mono(2-ethyl-5-carboxy-pentyl) terephthalate; MEHHP: mono(2-ethyl-5-hydroxyhexyl) phthalate; MEHP: mono(2-ethyl-hexyl) phthalate; MEOHP: mono(2-ethyl-5-oxohexyl) phthalate; MEP: mono-ethyl phthalate; MiBP: mono-iso-butyl phthalate; MiDP: mono-iso-decyl phthalate; MiNP: mono-iso-nonyl phthalate; MMP-9 or 2: matrix metalloproteinase; MMP: monomethyl phthalate; MeP: methylparaben; mTOR: mammalian target of rapamycin; MW: molecular weight; NF-κB: nuclear factor κ-light-chain-enhancer of activated B cells; NIS: sodium iodide symporter; NP: 4-nonylphenol; NRF-2: erythroid 2-related factor 2; OP: 4-tert octylphenol; PAX8: paired box 8; PE: polyethylene; Peg3: paternally expressed gene 3; PET: polyethylene terephthalate; PFR: placental-to-birthweight ratio; PHBA: para-hydroxybenzoic acid; PMCA1: plasma membrane calcium ATPase; PND: postnatal day; PrP: propylparaben; PPAR: peroxisome proliferator-activated receptors; PPPs: combination of phenols, parabens, and phthalates; PR: progesterone receptor; PVA: polyvinyl acetate; PVC: polyvinyl chloride; Rb: retinoblastoma; REACH: registration, evaluation, authorization, and restriction of chemicals; RORγ: retinoid Z receptor; SF-1: steroidogenic factor-1; SNAT-1 or 4: sodium-coupled neutral amino acid transporters; SRD5A2: steroid 5α-reductase 2; SULT1E1: estrogen sulfotransferase 1E1; T: testosterone; T4: thyroxine; TBAARS: thiobarbituric acid-reactive substances; TBBPA: tetrabromobisphenolA; TCS: triclosan; TET1: Tet methylcytosine dioxygenase 1; TGF-β: transforming growth factor β; THB: 2,3,4-trihydroxybenzophenone; TLR4: Toll-like receptor 4; TIMP-3: tissue inhibitor of metalloproteinase 3; TKs: toxicokinetics; TPO: thyroid peroxidase; TSH: thyroid-stimulating hormone; TTF-1: thyroid transcription factor 1; UE: molar urinary excretion of the measured compound; UGT1A1: UDP-glucuronosyltransferase 1A1; UVnorm: daily excreted urinary volume; Vss: steady-state volume of distribution; WAT: white adipose tissue; WC: waist circumference.
Footnotes
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Abstract
Phenols, parabens, and phthalates (PPPs) are suspected or known endocrine disruptors. They are used in consumer products that pregnant women and their progeny are exposed to daily through the placenta, which could affect offspring health. This review aims to compile data from cohort studies and in vitro and in vivo models to provide a summary regarding placental transfer, fetoplacental development, and the predisposition to adult diseases resulting from maternal exposure to PPPs during the gestational period. In humans, using the concentration of pollutants in maternal urine, and taking the offspring sex into account, positive or negative associations have been observed concerning placental or newborn weight, children’s BMI, blood pressure, gonadal function, or age at puberty. In animal models, without taking sex into account, alterations of placental structure and gene expression linked to hormones or DNA methylation were related to phenol exposure. At the postnatal stage, pollutants affect the bodyweight, the carbohydrate metabolism, the cardiovascular system, gonadal development, the age of puberty, sex/thyroid hormones, and gamete quality, but these effects depend on the age and sex. Future challenges will be to explore the effects of pollutants in mixtures using models and to identify the early signatures of in utero exposure capable of predicting the health trajectory of the offspring.
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Details
; Jovanovic, Nicolas 2
; Gayrard, Véronique 3 ; Mellouk, Namya 1 ; Marie-Noëlle Dieudonné 1 ; Picard-Hagen, Nicole 3 ; Flores-Sanabria, Maria-José 1
; Jammes, Hélène 1 ; Philippat, Claire 2 ; Couturier-Tarrade, Anne 1
1 Université Paris-Saclay, UVSQ, INRAE, BREED, 78350 Jouy-en-Josas, France; Ecole Nationale Vétérinaire d’Alfort, BREED, 94700 Maisons-Alfort, France
2 University Grenoble Alpes, Inserm U1209, CNRS UMR 5309, Team of Environmental Epidemiology Applied to Development and Respiratory Health, Institute for Advanced Biosciences, 38000 Grenoble, France
3 ToxAlim (Research Center in Food Toxicology), Université de Toulouse, INRAE, ENVT, INP-Purpan, UPS, 31062 Toulouse, France