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Per- and polyfluoroalkyl substances (PFAS) are a group of over 4700 heterogeneous compounds with amphipathic properties and exceptional stability to chemical and thermal degradation. The unique properties of PFAS compounds has been exploited for almost 60 years and has largely contributed to their wide applicability over a vast range of industrial, professional and non-professional uses. However, increasing evidence indicate that these compounds represent also a serious concern for both wildlife and human health as a result of their ubiquitous distribution, their extreme persistence and their bioaccumulative potential. In light of the adverse effects that have been already documented in biota and human populations or that might occur in absence of prompt interventions, the competent authorities in matter of health and environment protection, the industries as well as scientists are cooperating to identify the most appropriate regulatory measures, substitution plans and remediation technologies to mitigate PFAS impacts.
- PFAS
- PFOA
- PFOS
- human health
- ecosystem
- remediation technologies
1. Introduction
Per-and polyfluoroalkyl substances (PFAS) constitute an heterogeneous group of fluorinated synthetic compounds characterized by the presence of at least one perfluorinated methyl group (−CF3) or a perfluorinated methylene group (−CF2−), a variable number of carbon atoms, fluorination degree and presence of other chemical groups. PFAS are almost ubiquitous into the environment, mainly due to their wide dispersive use and applicability in a vast number of industrial sectors and consumer products [1,2]. Increasing concern for human health and wildlife ecology derives from the thermal and chemical stability of PFAS molecules and the multiple routes through which humans and biota can be exposed during their lifetime [3,4,5,6,7]. Of note, while the PFAS family has rapidly expanded into an impressive number of more than 4700 different substances including both the “legacy PFAS” (i.e., PFOS, PFOA) and the “emerging PFAS” (e.g., GenX) [8,9] producers, decision makers as well as researchers try to gain insights on their impact and to find the most appropriate measures to mitigate the potential risks associated with their exposure. Common features of PFAS are represented by their chemical stability which causes environmental persistence [10], their high mobility which confers them a long-range transport potential [11] causing their pervasive spreading even into remote regions (e.g., the Arctic’s or Antarctic’s) [12,13,14] and their tendency to bioaccumulate and biomagnify in biota through the contamination of the food chains [15,16,17,18,19]. The presence of some PFAS has been reported in the blood [20,21], milk [22,23], urine [24] tissues [25,26,27] and organs [28,29,30,31,32] of different human populations living in developed countries and has been associated to a number of adverse health effects. Similarly, relevant concentrations of PFAS have been detected in the air [33,34], groundwater [35,36], freshwater [17,37], marinewater [38,39], drinking water [40,41] and soil [42,43,44] potentially causing ecotoxic effects in the aquatic and terrestrial ecosystems at the trophic levels of primary producers, primary consumers and secondary consumers [45,46]. An additional layer of complexity is given by the coexistence of different mixtures of PFAS substances and other contaminants in the environmental media, for which quantitative risk assessment analysis and toxicologic/ecotoxicologic information is still scarce if not absent [47,48].
2. PFAS Human Exposure and the Potential Effects for Human Health
It is well recognized that multiple exposure pathways can link PFAS emission from the primary or secondary sources to human receptors represented by professional workers as well as general population. Some of the most relevant exposure routes include the inhalation of air and dust particulate, the ingestion of contaminated food and drinking water and the dermal adsorption [204]. Despite the presence of some gaps in our understanding, it is generally accepted that the dietary intake and the consumption of drinking water represent major pathways for the general population [205,206] while the relative contribute of inhalation and dermal contact is far more relevant in case of occupational exposure [207,208,209]. It is also believed that PFOS, PFOA, PFNA and PFHxS are the PFAS species that currently contribute most to human exposure, a reason for which provisional measures limiting their daily intake have been proposed in 2020 by the European Food Safety Authority (EFSA) [75]. Irrespectively to the specific exposure pathway through which they can get in contact with human targets, PFAS substances represent a serious concern for human health potentially inducing alterations in the development, lipid metabolism and endocrine system, cancerogenicity, immunotoxicity, hepatotoxicity and reprotoxicity. In terms of PFAS risk assessment, the EFSA CONTAM Panel was the first international scientific body to include the data from epidemiological studies to derive health-based guidance values for PFOA and PFOS in 2018. The critical effects of PFOS were identified as a rise in blood total cholesterol in adults and a reduction in antibody response to immunization in children. On the other hand, the key consequence of PFOA was a rise in blood total cholesterol. Reduced birth weight (for both chemicals) and an increased incidence of elevated blood levels of the liver enzyme alanine aminotransferase (ALT) (for PFOA) were also taken into account. Tolerable weekly intake (TWI) of 13 ng/kg body weight (b.w.) per week was defined for PFOS and 6 ng/kg b.w. per week for PFOA after benchmark modelling of serum levels of PFOS and PFOA and estimated the associated daily intakes. It was also observed that the exposure to a significant fraction of the population to both chemicals exceeded the proposed TWIs EFSA (2018) Risk to human health related to the presence of PFOS and PFOA in food [210]. The ATSDR (Agency for Toxic Substances and Disease Registry 2018) also included a decreased antibody response to vaccines (PFOA, PFOS, PFHxS, and PFDA) and increased risk of asthma diagnosis (PFOA) among the list of adverse health effects in PFAS-exposed humans (ATSDR Toxicological Profile for Perfluoroalkyls). Furthermore, the International Agency for Research on Cancer (IARC) classed PFOA as “possibly carcinogenic to humans” (Group 2B), based on limited evidence in humans that it might cause testicular and kidney cancer, as well as limited data in animal studies [211].
Potential of PFAS to cause a wide range of negative health impacts depends of various factors, such as the conditions of exposure (dose/concentration, duration, route of exposure, etc.) and characteristics associated with the exposed target (e.g., age, sex, ethnicity, health status, and genetic predisposition) [212]. The list of biological functions impacted by PFAS in females and males is rapidly expanding. Endocrine disruptive effects have been reported to affect fertility, body weight control, thyroid and mammary gland function. Developmental effects have been observed in children such as alterations in the behaviour or accelerated puberty but also in the new-borns such as decreased birth weight. Increased risk of kidney, prostate and testicular cancer has been associated with long-term exposure to PFAS in the general population alongside with disturbances in the cholesterol metabolism or reduced efficiency of the immune system against infections.
To illustrate this, in the following paragraphs we will describe more in detail some of the adverse effects caused by PFAS compounds, covering some of the recent in vitro, in vivo and epidemiological studies that have been published in the last ten years.
2.1. In Vitro Studies on PFAS Effects
Selected in vitro studies (published since 2010) exploring toxic effects of PFAS are summarised in Table 1. Majority of the presented in vitro studies investigated the impact of PFOA and/or PFOS, while the main observed effects were on thyroid [213,214,215,216] and hepatic cells [217,218,219,220,221]. In vitro exposure of thyroid cells to different PFAS was shown to have varied thyroid-disrupting effects. Conti et al. (2020) exposed thyroid follicular cells to 1–100 mM PFOS or PFOA, concluding that both substances acutely and reversibly inhibited iodide accumulation by FRTL-5 thyrocytes. Additionally, PFOS prevented sodium iodide symporter-mediated iodide uptake and reduced intracellular iodide concentration in iodide-containing cells. However, this substance did not affect iodide efflux from thyroid cells [213]. Furthermore, Song et al. (2012) documented decreased TPO activity in FTC-238/hrTPO/RSK008 cells after the exposure to different PFOS and PFOA concentrations [214]. At concentration of 105 nM PFOA/PFOS, a significant inhibition of cell proliferation was found in rat thyroid cell line-5 (FRTL-5), which mostly occurred due to the increased cell death. The results of this study also suggested that PFOA and PFOS enter thyroid cells by a gradient-based passive diffusion mechanism [215]. Croce et al. (2019) investigated the impact of different long-chain and short-chain PFAS, including PFOS, perfluorobutanesulfonic acid (PFBS), perfluorobutanoic acid (PFBA), perfluorophosphonic acid (PFPA) and perfluoropentanoic acid (PFPeA), on the same cell line (FRTL-5), at various concentrations (up to 100 μM). However, aside from PFOS (100 μM), neither long, nor short-chain PFCs impacted cell survival or interfered with cAMP synthesis. As a result, the authors came to the conclusion that short-chain PFCs had no acute cytotoxic effect on thyroid cells in vitro [216].
Table 1. Selected in vitro studies (published since 2010) exploring the toxicity of polyfluoroalkyl substances (PFAS).
| Cell Type | Substance | Treatment Concentration | Incubation Time | Effects | Ref. |
|---|---|---|---|---|---|
| thyroid follicular cells | PFOS PFOA |
PFOS or PFOA (1–100 mM) | Cytotoxicity: 1 h |
|
(Conti, Strazzeri, and Rhoden 2020) [213] |
| FTC-238/hrTPO/RSK008 cells | PFOS PFOA |
10−9, 10−8, 10−7, 10−6, 10−5, 10−4 M | / |
|
(Song et al., 2012) [214] |
| rat thyroid line-5 (FRTL-5) | PFOS PFOA |
1, 10, 102, 103, 104, and 105 nM | 72 h |
|
(Coperchini et al., 2015) [215] |
| rat thyroid line-5 (FRTL-5) | FOA, PFOS, perfluorobutanesulfonic acid (PFBS), perfluorobutanoic acid (PFBA), pentafluoropropionic anhydride (PFPA), perfluoropentanoic acid (PFPeA) |
0.0001; 0.001; 0.01; 0.1; 1; 100 μM | 24 h |
|
(Croce et al. 2019) [216] |
| Human hepatoma cell line (HepG2) | perfluorohexane sulfonate (PFHxS), perfluorooctane sulfonic acid (PFOS), perfluoroctanoic acid (PFOA), perfluorononanoate (PFNA), perfluorodecanoate (PFDA), perfluoroundecanoate (PFUnA), and perfluorododecanoate (PFDoA). |
2 × 10−7, 1 × 10−6, 2 × 10−6, 1 × 10−5, 2 × 10−5 M | 24 h |
|
(Wielsøe et al., 2015) [218] |
| Human Embryo Liver L-02 Cells | PFOS | 0, 50, 100, 150, 200, or 250 μmol/L | 24 or 48 h |
|
(Zeng et al., 2021) [219] |
| Human HepaRG liver cells | PFOA, PFOS, and perfluorononanoic acid (PFNA) | 6.25, 12.5, 25, 50, 100, 200, 400 μM | 6, 24, or 72 h |
|
(Louisse et al., 2020) [220] |
| HepaRG cell line | PFOS PFOA |
100, 250, 500, 750 μM PFOA 50, 100, 250, 500 μM PFOS |
/ |
|
Behr et al., 2021) [221] |
| Neurons | PFOS PFOA |
30–300 µM | 30 min |
|
(Liu et al., 2011) [222] |
| Primary rat cortical cultures and hiPSC-derived neuronal co-cultures | PFOS PFOA |
0.01, 0.1, 1, 10, 100 µM | / |
|
(Tukker et al., 2020) [223] |
| Rat primary hippocampal neurons and astrocytes |
PFOS | 25, 50, 75, 100, 125 μM for neurons 15, 25, 50, 75, 100 μM for astrocytes |
24 h |
|
(Li et al., 2017) [224] |
| primary rat embryonic neural stem cells (NSCs) | PFOS | 12.5–100 nM | 48 h |
|
(Wan Ibrahim et al., 2013) [225] |
| rat primary neurons and neural stem cells (NSC) | PFOS PFOA |
1–250 μM | 24 h |
|
(Pierozan and Karlsson 2021) [226] |
| fetal rat testes or seminiferous tubule segments (stage VII-VIII) of adult rats |
PFOA | 0–100 μg/mL | 24 h |
|
(Eggert et al., 2019) [227] |
| human cell lines such as MCF-7, H295R, LNCaP and MDA-kb2 | PFOA, PFOS, and of six substitutes including perfluorohexanesulfonic acid (PFHxS), perfluorobutanesulfonic acid (PFBS), perfluorohexanoic acid (PFHxA), perfluorobutanoic acid (PFBA), ammonium perfluoro(2-methyl-3-oxahexanoate) (PMOH), and 3H-perfluoro-3-[(3-methoxypropoxy) propanoic acid] (PMPP) |
various concentrations | 24 h when cytotoxicity was assayed in HEK293T, LNCaP or MDA-kb2 cells, for 6 d in MCF-7 cells and for 48 h in H295R cells |
|
(Behr et al.,2018) [228] |
Some of the in vitro studies explored liver toxicity of the PFAS by investigating the occurrence of the oxidative stress on different hepatic cell lines [218,219], as well as apoptosis/autophagy [219] and cholesterol/bile acids level [221]. The impact of PFOS, PFOA, PFHxS, PFNA, PFDA, PFUnA, and PFDoA on the same cell line (HepG2) has been tested. Here, PFHxS, PFOA, PFOS, and PFNA caused a dose-dependent increase in DNA. With the exception of PFDoA, all other PFAS elevated ROS formation. Cells exposed to PFOA had significantly lower total antioxidant capacity (TAC) compared to the control, but PFOS and PFDoA had a non-significant trend in TAC decrease and an increasing tendency for PFHxS, PFNA, and PFUnA [218]. Other studies also explored the link between PFOA/PFOS and oxidative stress/apoptosis in different liver cells. Zeng et al. (2021) tested the influence of PFOS on human embryo liver L-02 cells. Reduced cell activity was observed, as well as increased ROS levels in a concentration-dependent manner. Impaired mitochondrial membrane potential (MMP), as well as elevated autophagy and apoptosis were observed, together with the increased expression of Bax, cleaved-caspase-3, and LC3-II. The authors concluded that ROS-dependent autophagy might be the cause of PFOS-induced apoptosis in L-02 cells [219]. Other mechanisms underlying PFAS-induced liver injury were also investigated. Human HepaRG liver cells were treated to PFOA, PFOS, and (PFNA at various doses. All PFAS increased cellular triglyceride levels while having no effect on cholesterol levels. In HepaRG cells, PFOA, PFOS, and PFNA enhanced triglyceride levels while inhibiting cholesterogenic gene expression. The authors noted that PFAS-induced endoplasmic reticulum stress might be an essential mechanism underpinning some of the harmful effects caused by these compounds [220]. Similarly, Behr et al. (2021) investigated the effects of PFOS/PFAS exposure and found that cholesterol levels in the same cell line (HepaRG) were not affected by neither PFOA nor PFOS. However, in this study, both substances strongly decreased synthesis of a number of bile acids. Moreover, the expression of numerous genes whose products are involved in synthesis, metabolism and transport of cholesterol and bile acids was strongly affected by PFOA and PFOS at concentrations above 10 µM. Indeed, PFOS and PFOA led to a strong decrease of CYP7A1, the key enzyme catalysing the rate-limiting step in the synthesis of bile acids from cholesterol, both at the protein and the mRNA level. Both substances led to a dilatation of bile canaliculi [221].
Oxidative stress and apoptosis/autophagy were also investigated in the light of PFAS-induced neurotoxicity. Li et al. (2017) exposed rat primary hippocampal neurons and astrocytes to PFOS, which led to redox imbalance, increased apoptosis and abnormal autophagy. In astrocytes, PFOS altered extracellular glutamate and glutamine concentrations, decreased glutamine synthase activity and impaired gene expression of glutamine synthase as well as glutamate and glutamine transporters [224]. Other PFAS-linked neurotoxic mechanisms included interferences at the gene expression level and calcium homeostasis. Wan Ibrahim et al. (2013) investigated the effect of PFOS on primary rat embryonic neural stem cells (NSCs) and noted an increase in neuronal differentiation. Upregulation of PPARγ was also observed, with no changes in PPARα or PPARδ genes. Additionally, PFOS upregulated mitochondrial uncoupling protein 2 (UCP2) and induced Ca2+ activity [225]. Both PFOS and PFOA were found to accumulate in cultured neurons and elevate calcium concentrations via release of intracellular calcium stores. 1,4,5-trisphosphate receptors (IP3Rs) and ryanodine receptors (RyRs) were found to take part in PFOS or PFOA induced calcium release, which caused a perturbation of calcium homeostasis [222]. PFOS and PFOA also inhibited the GABA-evoked current in primary rat cortical cultures and acted as non-competitive human GABA-A receptor antagonists. Network activity of rat primary cortical cultures increased following exposure to PFOS [223]. Pierozan and Karlsson (2021) studied the effects of PFOS and PFOA on primary neurons and NSC from rats. There were no impacts on cell viability or proliferation in primary neurons. At the lowest studied concentration (1–100 μM), PFOS exposure boosted NSC proliferation, while PFOS and PFOA altered the morphology of NSC-derived neurons. The neurite network was similarly impacted by 1 and 10 μM PFOA exposure, resulting in an increase in the number of processes and branches per cell. The authors concluded that NSC, which mimics the embryonic brain, are more vulnerable to PFOS and PFOA exposure than neurons [226].
PFAS effects on reproductive function were also investigated in in vitro studies. Eggert et al. (2019) assessed the effects of 0–100 μg/mL PFOA on foetal rat testes and adult rat seminiferous tubule segments. The findings revealed lower levels of cAMP, progesterone, testosterone, and StAR expression. The number of diploid, proliferative, meiotic, and G2/M-phase cells in adult rat testis was drastically reduced by PFOA. By contrast, PFOA had no effect on foetal, proliferating, or adult rat Sertoli cells. Foetal Leydig cells, on the other hand, showed an increased susceptibility to apoptosis [227]. The exposure of human cell lines such as MCF-7, H295R, LNCaP and MDA-kb2PFOA to PFOS and six substitutes including PFHxS, PFBS, PFHxA, PFBA, PMOH and ADONA has revealed that PFOA, PFOS and PMOH enhanced 17β-estradiol-stimulated estrogen receptor β activity, and PFOS, PMOH, PFHxA and PFBA enhanced dihydrotestosterone-stimulated androgen receptor activity. PFOA and PFOS slightly enhanced estrone secretion, and progesterone secretion was marginally increased by PFOA. All these effects were only observed at concentrations above 10 μM [228].
2.2. In Vivo Studies on PFAS Effects
PFAS exposure has mostly been linked to adverse outcomes in mice and rats, while the majority of studies explored the toxicity of PFOS, PFOA and PFHxS. Selected in vivo studies (published since 2010) investigating toxic effects of PFAS are presented in Table 2.
Table 2. Selected in vivo studies (published since 2010) exploring the toxicity of polyfluoroalkyl substances (PFAS).
| Species | Substance | Dose and Route of Exposure | Exposure Time | Effects | Ref. |
|---|---|---|---|---|---|
| Rats | PFOS | 20 or 100 ppm, dietary exposure | 7 days |
|
(Elcombe et al., 2012) [229] |
| Mice | PFOS | 10 mg PFOS/kg b.w./day), oral gavage | 14 days |
|
(D. Li et al., 2021) [230] |
| Mice | PFOS | 100 μg/kg b.w./day and 1000 μg/kg b.w./day, oral gavage | 2 months |
|
(X. Li et al., 2021) [231] |
| Rats | PFOA | 5 mg/kg b.w./day, oral gavage | 28 days |
|
(Owumi, Bello, and Oyelere 2021) [232] |
| Mice | PFOA | 1, 5, 10, or 20 mg/kg/day, oral gavage | 10 days |
|
(Rashid et al., 2020) [233] |
| Mice | PFHxS | Up to 3 mg/kg b.w./day, oral gavage | Administered before mating, for at least 42 days in F0 males, and for F0 females, through gestation and lactation. F1 pups-directly for 14 days after weaning |
|
(Chang et al., 2018) [234] |
| Rats | PFHxS | 0.05, 5 or 25 mg/kg b.w./day, oral gavage |
From gestation day 7 through to postnatal day 22 |
|
(Ramhøj et al., 2020) [235] |
| Mice | 6.1, and 9.1 mg/kg b.w., oral gavage | Neonatal exposure from postnatal day 10 |
|
(Sim and Lee, 2022) [236] |
A link between PFOS exposure and hepatic lipid metabolism was found in vivo. After the exposure of C57BL/6 mice to 10 mg PFOS/kg b.w./day by oral gavage for 14 days, 241 proteins involved in lipid and xenobiotic metabolism in liver were found to be dysregulated. 16 overexpressed glycoproteins were associated with neutrophil degranulation, cellular responses to stress, and protein processing in the endoplasmic reticulum (ER) [230]. Li et al. (2021) also documented the ability of PFOS to accumulate in the liver of female BALB/c mice after the exposure to 100 μg/kg b.w./day and 1000 μg/kg b.w./day by oral gavage for 2 months. PFOS accumulated in the lungs, kidneys, spleen, heart and brain as well, causing damage in the liver and in the marginal area of the heart. PFOS mainly affected glycerophospholipid metabolism and sphingolipid metabolism in liver. These authors suggested that the upregulated ceramide and lysophosphatidylcholine (LPC) might lead to liver cell apoptosis, while decrease in liver triglyceride (TG) content might result in insufficient energy and cause liver morphological damage [231]. After the dietary exposure of rats to 20 or 100 ppm PFOS for 7 days, Elcombe et al. (2012) have also noted alterations in various liver parameters (e.g., increased liver weight; decreased plasma cholesterol, alanine aminotransferase, and triglycerides; increased hepatocellular cytosolic CYP450 concentration; increased liver activity of acyl CoA oxidase, CYP4A, CYP2B, and CYP3A; increased liver proliferative index and decreased liver apoptotic index). However, in the aforementioned study, thyroid parameters (histology, apoptosis, and proliferation) were not unaffected [236].
Owumi et al. (2021) investigated hepatic and kidney function of PFOA. After the exposure of rats to PFOA (5 mg/kg b.w./day) for 28 days, the authors noted an increase in hepatic (GGT, ALT, AST and ALP) and renal function (urea and creatinine) as well as biomarkers of toxicity, paralleled by a decrease in the activity of the enzymatic antioxidants (CAT, GPx, SOD) in liver and kidney tissue. In this study, a significant increase in lipid peroxidation and pro-inflammatory cytokine IL-1β in rats’ liver and kidney occurred, while a decrease of the anti-inflammatory cytokine, IL-10 was also observed [232]. Similarly, renal damage was found in mice after the exposure to PFOA. Rashid et al. (2020) noted an increase in Dnmt1 with decreased Rasal1 expression at higher levels of PFOA exposure. Rasal1 hypermethylation (an early indicator of fibroblast activation in kidney) was also observed, followed by the increase in Hdac1, 3 and 4, class I & II HDACs which are known to be critically altered in some renal diseases. Furthermore, mRNA levels of TGF-β and α-SMA were significantly increased [233].
Health effects of PFHxS have also been investigated in animal studies. After the oral exposure of F0 and F1 CD-1 mice to 3 mg PFHxS/kg b.w/day, equivocal decrease in live litter size at 1 and 3 mg/kg b.w./day was noted, as well as adaptive hepatocellular hypertrophy, concomitant decreased serum cholesterol and increased alkaline phosphatase [234]. After treating rat with 0.05, 5 or 25 mg/kg b.w./day PFHxS from gestation day 7 onward to postnatal day 22, Ramhøj et al. (2020) observed a dose-dependent decrease in thyroid hormone levels in both dams and offspring, while TSH levels, weight, histology, or expression of marker genes of the thyroid gland were unaffected [235]. Other neurotoxic effects of PFHxS have been investigated as well. Sim and Lee (2022) have found that PFHxS causes long-term developmental neurotoxicity as well as downregulation of GAP-43 and CaMKII via the NMDA receptor-mediated PKCs (and)-ERK/AMPK pathways in mice after the neonatal exposure, together with significant memory impairment in adult mice [236].
2.3. Human Studies on PFAS Effects
Selected human studies (published since 2010) investigating toxic effects of PFAS are presented in Table 3. Majority of the human studies explored the linkage between PFAS concentration and lipid status, mainly cholesterol level [237,238], while a study was also conducted to assess the connection between PFAS and cholesterol at the gene expression level [239]. Eriksen et al. (2013) discovered substantial positive relationships between PFOS, PFAS, and total cholesterol in 753 individuals, while sex and prevalence of diabetes were suggested to influence the connection between these two substances and cholesterol [237]. Fletcher et al. (2013) observed an inverse relationship between serum PFOA levels and the expression level of genes involved in cholesterol transport in whole blood (NR1H2, NPC1 and ABCG1). A positive correlation was found between PFOS and a transcript involved in cholesterol mobilization (NCEH1), while a negative relationship was seen between PFOS and a transcript involved in cholesterol transport (NCEH2). Sex-specific effects were also noticed in this study [239]. On the other hand, in a study involving 815 participants ≤18 years of age, Geiger et al. (2014) found that serum PFOA and PFOS were related with high total cholesterol and LDL-C levels, regardless of age, gender, race-ethnicity, body mass index, yearly family income, physical activity, or serum cotinine levels. PFOA and PFOS were not shown to be substantially linked with aberrant HDL-C and triglyceride levels [238].
Table 3. Selected human studies (published since 2010) exploring the toxicity of polyfluoroalkyl substances (PFAS).
| Substance | Population | Measured Parameters | Results | Ref. |
|---|---|---|---|---|
| PFOS PFOA |
middle-aged Danish population; 753 individuals (663 men and 90 women), 50–65 years of age, nested within a Danish cohort of 57,053 participants |
serum levels of total cholesterol |
|
(Eriksen et al., 2013) [237] |
| PFOS PFOA |
815 participants ≤18 years of age from the National Health and Nutrition Examination Survey 1999–2008 |
dyslipidemia: total cholesterol >170 mg/dL, low-density lipoprotein cholesterol (LDL-C) >110 mg/dL, high-density lipoprotein cholesterol (HDL-C) <40 mg/dL or triglycerides >150 mg/dL. |
|
(Geiger et al., 2014) [238] |
| PFOS PFOA |
290 individuals (144 men + 146 women) exposed to background levels of PFOS and elevated concentrations of PFOA through drinking water, aged between 20 and 60 years |
expression of genes involved in cholesterol metabolism |
|
(Fletcher et al., 2013) [239] |
| PFOA PFOS PFHxS PFNA PFDA |
2883 participants, (1801 non-obese and 1082 obese), aged more than or equal to 20 years old |
liver function parameters: AST, ALT, GGT, ALP, and total bilirubin (TB) |
|
(Jain and Ducatman 2019) [240] |
| 14 PFCs | Healthy men from the general population, median age of 19 years | total testosterone (T), estradiol (E), sex hormone-binding globulin (SHBG), luteinizing hormone (LH), follicle-stimulating hormone (FSH) and inhibin-B and Semen samples analysis |
|
(Joensen et al., 2013) [241] |
| PFOA PFOS PFHxS PFNA |
1682 males and females 12 to 80 years of age |
testosterone (T), thyroid stimulating hormone (TSH), and free and total triiodothyronine (FT3, TT3) and thyroxine (FT4, TT4) |
|
(Lewis, Johns, and Meeker 2015) [242] |
| PFOS PFOA |
3076 boys and 2931 girls aged 8–18 years |
subjects were classified as having reached puberty based on either hormone levels (total >50 ng/dL and free >5 pg/mL testosterone in boys and estradiol >20 pg/mL in girls) or onset of menarche |
|
(Lopez-Espinosa et al., 2011) [243] |
| PFOS PFOA PFNA |
2292 children (6–9 years of age) | estradiol, total testosterone, and IGF-1 |
|
(Lopez-Espinosa et al., 2016) [244] |
| PFOS PFOA |
424 mother-infant pairs | estrone (E1), b-estradiol (E2), and estriol (E3), infants: head circumference, body weight, body length |
|
(Wang et al., 2019) [245] |
| PFOS PFOA |
47,092 adults |
alanine transaminase (ALT), γ-glutamyltransferase (GGT), direct bilirubin |
|
(Gallo et al., 2012) [246] |
| PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFHxS PFOSA |
1002 individuals from Sweden (50% women) at ages 70, 75 and 80 | bilirubin and hepatic enzymes alanine aminotransferase (ALT), alkaline phosphatase (ALP), and γ-glutamyltransferase (GGT) |
|
(Salihovic et al., 2018) [30] |
| PFOS PFOA PFHxS |
3297 participants from Ronneby, a municipality with drinking water highly contaminated by PFAS (exposed group) | thyroid hormone levels, with adjustments for age, sex and BMI |
|
(Y. Li et al., 2021) [247] |
| PFOA PFOS |
101 healthy 1-year-old children | Antibodies against haemophilus infuenza type b, tetanus and diphtheria, interferon gamma, cholesterol |
|
(Abraham et al., 2020) [248] |
| PFOA PFOS |
1146 children | serum concentrations of specific IgG antibodies against tetanus and diphtheria at ages 5 and 7 |
|
(Budtz-Jørgensenet al., 2018) [249] |
| PFHxS, PFOS, PFOA, PFDA, PFNA | 275 males and 349 females participated in clinical examinations and provided blood samples at ages 18 months and 5 years | serum concentrations of antibodies against tetanus and diphtheria vaccines determined at age 5 |
|
(Grandjean et al., 2017) [250] |
| PFHxS, PFOA, PFOS, PFNA, PFDA. | 516 subjects | PFAS serum concentrations and concentration of antibodies against diphtheria and tetanus |
|
(Grandjean et al., 2017) [251] |
Other studies explored the link between PFAS concentration and different hormones, such as thyroid [242,247] and sex hormones [245,247,248], as well as development [243,245]. By assessing the connection between the levels of 14 PFAS in healthy men from the general population and different sex hormones and semen sample quality, Joensen et al. (2013) found that only PFOS levels were negatively associated with testosterone, calculated free testosterone (FT), free androgen index (FAI) and ratios of T/LH, FAI/LH and FT/LH. Other PFAS were found at lower levels than PFOS and did not exhibit the same associations [241]. Also, after measuring PFAS levels in 1682 males and females 12 to 80 years of age, Lewis et al. (2015) found no significant relationships between any of the PFAS and testosterone. PFAS were suggested to be associated with increases in FT3, TT3, and FT4 among adult females. The authors concluded that, during the adolescence, PFAS may be related to increases in TSH among males and decreases in TSH among females [242], suggesting sex-specific effects. In contrast, Li et al. (2021) discovered no associations between PFAS and thyroid hormones in adults and seniors in 3297 participants from Ronneby, a municipality with highly contaminated drinking water by PFAS (exposed group), with the exception of a positive association between PFAS and fT4 in males over 50. Thyroid hormone levels were observed to be higher in Ronneby preteen children compared to the control group. Weak evidence of a link between increasing PFAS levels and lower fT3 in preteen boys and lower TSH in adolescent men was found [247].
Lopez-Espinosa et al. (2011) aimed to investigate whether PFOS and PFOA were linked to the markers of sexual development. Their study included 3076 boys and 2931 girls aged 8–18 years. For boys, there was a link between increased PFOS and a lower chance of reaching puberty. Higher PFOA or PFOS concentrations in girls were related to a lower risk of post menarche [243]. The same group of researchers examined the link between PFAS levels and estradiol, total testosterone, and IGF-1 in 2292 children. In boys, PFOA concentrations were substantially related to testosterone levels; PFOS concentrations were related to estradiol, testosterone, and IGF-1, while PFNA concentrations were linked to IGF-1. Significant linkage was discovered in girls between PFOS and testosterone and IGF-1, as well as PFNA and IGF-1 [244]. Furthermore, Wang et al. (2019) concluded that PFAS may affect estrogen homeostasis and foetal growth during pregnancy, and that estrogens may mediate the relationship between PFAS exposure and foetal growth after examining 424 mother-infant pairs [245].
Some of the studies also explored the linkage between the PFAS exposure and liver function [31,251,252]. In 47,092 adult participants, Gallo et al. (2012) found a positive association between PFOA and PFOS concentrations and serum ALT level. On the other hand, the relationship with bilirubin appeared to increase at low levels of PFOA and decrease at higher levels [246]. In 1002 individuals from Sweden, Salihovic et al. (2018) have also found a positive association of PFHpA, PFOA, PFNA, and PFOS concentrations and ALT activity, but also positive associations of PFHpA, PFOA, PFNA, PFDA, and PFUnDA with ALP. These authors noted that the changes of investigated PFAS concentrations were positively associated with gamma glutamyl transferase (GGT) levels and inversely associated with the changes in circulating bilirubin [30]. On the other hand, in 2883 participants, (1801 non-obese and 1082 obese), Jain and Ducatman investigated the connection between liver function alterations and various PFAS. They concluded that connections might only be observed in the obese participants: alanine aminotransferase (ALT) was positively associated with PFOA, PFHxS, and PFNA. On the other hand, PFOA and PFNA were associated with GGT [240]. Epidemiological studies revealed a connection between PFAS and decrease in vaccination antibody production in early infants and children, especially having in mind that, if breastfed, they have a relatively high exposure and may be more susceptible as their immune system develops. Abraham et al. (2020) found significant associations between the concentration of PFOA, but not PFOS, and adjusted levels of vaccine antibodies against Haemophilus influenza type b, tetanus and diphtheria for which no observed adverse effect concentrations (NOAECs) were 12.2, 16.9 and 16.2 µg/L, respectively. Furthermore, PFOA levels were shown to be inversely related to the interferon gamma (IFN-γ) production of ex-vivo lymphocytes after stimulation with tetanus and diphtheria toxoid [248]. Furthermore, Budtz-Jorgensen E and Grandjean P (2018) found an approximate BMDL of 1 ng/mL serum for both PFOS and PFOA for the serum concentrations of specific IgG antibodies against tetanus and diphtheria at ages 5 and 7 as outcome parameters. These authors proposed the reference concentration of about 0.1 ng/mL as the serum-based target, a level which is below the most reported human serum-PFAS concentrations [249]. Grandjean et al. (2017) discovered that prenatal exposure to PFAS had an inverse relationship with antibody concentrations five years later, while concentrations measured at 3 and 6 months of age had the highest inverse relationships with antibody concentrations at 5 years of age, especially for tetanus [250]. The same authors have found that diphtheria antibody concentrations dropped at higher PFAS concentrations at 13 and 7 years after booster vaccinations at 5 years of age; the correlations were statistically significant for PFDA at 7 years and PFOA at 13 years, implying a 25% decrease for each doubling of exposure [251].
This entry is adapted from the peer-reviewed paper 10.3390/toxics10020044
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