Per- and Polyfluorinated Alkyl Substances (PFAS) are synthetic chemicals that contain at least one perfluoroalkyl group (Birru et al., 2021). Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) are the most common PFAS detected in the environment due to their widespread use in manufacturing and chemical stability (Huang and Jaffé, 2019). Both PFOS and PFOA are considered long-chain PFAS because they possess an eight-carbon backbone; their functional groups are sulfonate and carboxylate, respectively (Tsuda, 2016). They possess both hydrophobic and oleophobic properties, along with other chemical characteristics making them useful in many consumer goods (Birru et al., 2021). Their carbon-fluorine bonds, make them resistant to degradation and allow them to persist in the environment and bioaccumulate within living organisms (Zeng et al., 2019; Neagu et al., 2021).

After more than 50 years of production, PFOS and PFOA were widely discovered in humans and the environment, and a subsequent phasing out of production by the 3M company began in 2000 (Giesy and Kannan, 2001; Olsen et al., 2017). Subsequently, the United States Environmental Protection Agency procured an agreement with other major manufacturing companies to reduce long-chain PFAS emissions; other global environmental regulations for PFAS emerged as well (Olsen et al., 2017). The phasing out of these compounds resulted in a decline in human blood serum concentrations of PFOS and PFOA between the years of 2000 and 2015 (Olsen et al., 2017; Wang et al., 2017). Despite the reduction in production of long-chain PFAS, the compounds persist in the environment and in humans and are a continuing concern due to their widespread environmental persistence, distribution, potential toxicological effects, potential to bioaccumulate (Huang and Jaffé, 2019).

Ingestion is considered the largest exposure route to humans (rather than dermal and inhalation). Tap water is a major PFAS exposure route where most average PFAS concentrations are in the low ng/L range (Sinclair et al., 2020). Regarding PFOA and PFOS, the USEPA has issued a Health Advisory Level of 70 ppt and is in the process of making a final regulatory determination (USEPA, 2021). Seafood is a significant PFAS exposure route while exposure from consumer goods, food packaging, and indoor environments are uncertain (Sunderland et al., 2019).

With the phasing out of long-chain PFAS in the early 2000s, short-chain alternatives (i.e., those with carbon backbones of <7 carbons) began to take their place in industry and the environment (Brendel et al., 2018). These alternatives share similar structures to long-chain PFAS (i.e., heavily fluorinated carbon chains), however, one or more alkylether group is inserted into a shorter fluoroalkyl chain. Short-chain PFAS are also products of long-chain degradation (Li et al., 2020). Short-chain replacements GenX, ADONA, and F53B are widely used. GenX is a trade name for a short-chain PFAS processing technology, of which the hexafluoropropylene oxide (HFPO) dimer acid and its ammonium salt are the major constituents. ADONA (dodecafluoro-3H-4,8-dioxanonanoate) is generally used as a substitute for PFOA in fluoropolymer production (Fromme et al., 2017; Munoz et al., 2019). Chlorinated polyfluoroalkyl ether sulfonate, or F53B, has been produced as an alternative to PFOS and has been adopted as a mist suppressant by several electroplating companies (Du et al., 2016; Munoz et al., 2019; Shi et al., 2019).

Water is the major environmental sink for long- and short-chain PFAS. Bai and Son (2021) quantified concentrations of 17 PFAS in the Las Vegas and Reno watersheds (Nevada, USA), and compared them to data on aquatic PFAS concentrations worldwide. Overall, Bai and Son (2021) found that concentrations of individual PFAS ranged from non-detectable to low ppt (maximum of 74.47 ng/L) which was similar to studies and findings in watersheds of the United States (Nakayama et al., 2007; Zhang et al., 2016), Europe (Moller et al., 2010), Uganda (Dalahmeh et al., 2018), China (So et al., 2007; Zhao et al., 2013; Wang et al., 2019), and Australia (Clara et al., 2009). In surface waters where sewage effluent is the primary input, maximum PFAS concentrations were 207.59 ng/L; groundwater contaminated with landfill leachate maximum PFAS concentrations were 5,200 ng/L (Zhou W. et al., 2017). PFAS have been detected in the tissues of many wild species, including, but not limited to, bald eagles, albatrosses, polar bears, seals, dolphins, alligators, squid, and many fish species (Giesy and Kannan, 2001; McCarthy et al., 2017).

Although atmospheric deposition is a possibility, the main source of GenX and ADONA in surface water is manufacturer’s wastewater (Hopkins et al., 2018). F53B is primarily released into waterways via chrome plating industries, finding its way into nearby irrigation systems and ground water. The shorter length and alkylether substitution of short-chain PFAS contribute to a greater water solubility. GenX, when in water, loses an ammonium group, creating an anion that is the same as the HFPO-DA anion in water (Hopkins et al., 2018). This anion has a precursor that travels via air, the C3 dimer acid fluoride (Hopkins et al., 2018). Once the C3 dimer acid fluoride interacts with water, it creates the same anion which readily migrates to groundwater, contributing to private well contamination (Hopkins et al., 2018). The limited data indicate that most short-chain PFAS are more persistent and widespread than long-chain PFAS in the aquatic environment (Olsen et al., 2017; Wang et al., 2017; Brendel et al., 2018; Li et al., 2020). GenX chemicals are present in multiple aquatic environments including surface water, groundwater, and in drinking water and rainwater (USEPA, 2018). Aside from factory wastewater effluent or points very close to industry outfalls, concentrations of GenX, ADONA, and F53B are found in surface waters in the low ng/L concentrations (Pan et al., 2018; Munoz et al., 2019; Gebbink and van Leeuwen, 2020; Bai and Son, 2021). Both GenX and F53B were found in higher concentrations than long-chain PFAS in surface water samples collected downstream from a PFAS manufacturer (Gebbink et al., 2017; Ateia et al., 2019). And although short-chain PFAS have been quantified in drinking water and sediment worldwide, most studies did not focus on quantitation of GenX, ADONA, and F53B and thus, data in the aquatic environment are lacking (Ateia et al., 2019).

Bai and Son (2021) found that short-chain PFAS were more likely to partition in the water column, while long-chain PFAS most likely to partition in sediment. This is similar to the finding of Zhao et al. (2016) who estimated that 88.8% of total PFAS in the water column are short-chain. These estimations correspond to what would be expected based on the greater water solubility of short-chain PFAS.

It is important to note that the studies that concluded “relatively low” concentrations of PFAS in the water column or sediment are often describing the concentration of individual molecules. While environmental regulations limit concentrations of individual PFAS, the sum of the thousands of PFAS need to be considered for estimations of exposure and effects. Indeed, even if toxicity is considered to be simply additive, estimations of risk must include total PFAS exposure.

Researchers have utilized many different model organisms in order to better understand the reproductive effects of PFOS on living organisms. Shi et al. (2008) observed zebrafish embryos exposed 4 h post-fertilization to varying concentrations of PFOS between 0.1 and 5 mg/L, and observed significant changes in embryo development. Noted changes included delayed hatching, malformation, and depressed heart rates, which led to reduced survival rates of fry exposed to 1 mg/L PFOS for 21 days before hatching (Shi et al., 2008). Han and Fang conducted a study to observe the effects of PFOS exposure on swordtail fish; they found a significant decrease in survival rate of offspring at higher levels of PFOS exposure in these swordtail fish (Han and Fang, 2010). PFOS exposure has been associated with inhibition of sperm production (Qu et al., 2016). In female mice, PFOS exposure is associated with ovulation reduction (Wang et al., 2018). Exposure of pregnant mice to PFOS resulted in neonatal mortality as a result of severe lung collapse and intracranial blood vessel dilation (Yahia et al., 2010).

While PFOA is similar to PFOS in structure, PFOS is generally considered to be more toxic (Zheng et al., 2012). Similar research has been conducted on PFOA as that on PFOS in order to understand the reproductive effects of the compound. A study was conducted to examine the effects of PFOA on development of zebrafish and found that PFOA is acutely toxic to zebrafish embryos (LC₅₀ = 262 mg/L), similar to PFOS (Zheng et al., 2012). At higher concentrations, edema, delayed hatching, and spinal malformations were observed, and nearly 100% mortality was observed when the embryos were exposed to 270 mg/L (Zheng et al., 2012). A meta-analysis was conducted to determine the effects of PFOA on the reproductive system of male rats and found significant association between higher PFOA levels and reproductive toxicity (Wang et al., 2021). Toxicological effects of increased serum PFOA levels included a decrease in serum testosterone levels and a decrease in weight of reproductive organs (testicle and epididymis) (Wang et al., 2021). The exposure of PFOA to pregnant mice resulted in reduced fetal body weight and delayed bone formation, and a 100% mortality rate of pups after exposure at 10 mg/kg was observed (Yahia et al., 2010). PFOA increased mortality of the pups, similarly to PFOS, but no intracranial blood vessel dilation was observed, indicating a different cause of death by PFOA (Yahia et al., 2010).

PFAS interact directly with estrogen receptor α (ERα) in trout liver in vivo, although they have a weak affinity and they are noted to induce an estrogenic effect (Benninghoff et al., 2010). In male mice, PFOS exposure was observed to down-regulate ERα and up-regulate estrogen receptor β, which increased cell apoptosis and decreased cell proliferation in the testes (Qu et al., 2016). In female mice, PFOS was shown to suppress ERα receptors, which led to induced prolongation of diestrus and ovulation reduction (Wang et al., 2018). Accordingly, Jensen and Leffers (2008) found that male rats exposed to PFOA developed Leydig cell hyperplasia eventually resulting in adenomas (Jensen and Leffers, 2008). A decrease in testosterone production was also reported in these rats followed by an increase in estradiol levels.

Reproductive functions are dependent on peroxisome proliferator-activated receptors (PPARs). These receptors are susceptible to PFAS which are known to affect gene transcription through the activation of PPARs (Vitti et al., 2016). When mice were treated with PFOA, subsequent changes in PPARs were associated with delayed mammary gland development (Zhoa et al., 2012).

Three recent studies examined reproductive effects of GenX in mice (Blake et al., 2020) and rats (Conley et al., 2019; Conley et al., 2021). Mice treated with GenX had a greater incidence of placental abnormalities while rats showed a greater incidence of reduced maternal thyroid hormone levels and elevated levels of PPAR-regulated gene expression in both maternal and fetal livers (as with PFOA described above). GenX is a developmental toxicant in the rat; those dosed from gestational day 8 through post-natal day 2 (from 1 to 125 mg/kg/day), had increased neonatal mortality and reduced birth weight (Conley et al., 2021).

In zebra fish, F53B was associated with an increase in birth defects, delayed hatchings, and decreased survival rates in embryos (Shi et al., 2017). Conversely, Gaballah et al. (2020) concluded that GenX (and ADONA) was not associated with developmental toxicity in zebrafish.

There is conflicting evidence regarding whether or not PFAS interact with estrogen receptors in humans. Multiple studies have concluded that PFAS promote activity of estrogen receptors in the presence of estradiol, but do not directly interact with them (Kang et al., 2016; Behr et al., 2018). Another study found that PFAS do not have any effect on ER, while another found that PFAS bind directly to ER without the presence of estradiol (Kjeldsen and Bonefeld-Jørgensen, 2013; Yao et al., 2014). Despite being heavily researched, a lack of understanding of the mode of action of these pollutants persists (Cousins et al., 2020).

Zhang et al. (2015) exposed human syncytiotrophoblasts isolated from term placenta to PFOS. Dose-dependent reductions of human chorionic gonadotropin (hCG), estradiol, and progesterone resulted. After blocking the apoptotic route, the authors concluded that PFOS exerted its effects by inducing apoptosis due (Zhang et al., 2015).

Multiple epidemiological studies have been conducted to determine the associated effects of PFAS on humans. Johnson et al. (2014) found an association between lower birth weights and exposure to PFOA, noting an 18.9 g decrease in birth weight for every ng/mL increase in serum PFOA level. As described in rat models above (Jensen and Leffers, 2008), Leydig cell hyperplasia is connected with lower testosterone levels in men, indicating a possible mode of action for PFOA to induce testicular cancer (Tarapore and Ouyang, 2021).

As also described above, the effects of PFAS on human reproductive hormones is uncertain. One epidemiological study found that an increase in human serum PFAS levels was negatively associated with serum levels of reproductive hormones, including testosterone (Tsai et al., 2015). However, another study found a positive correlation between increased PFAS levels and levels of testosterone and estradiol, while yet another study found no correlation between PFAS and sex hormone levels (Olsen et al., 1998; Sakr et al., 2007). PFAS have also been shown to compete with thyroxine to bind to transthyretin (Kar et al., 2017). Both molecules play a major role in thyroid function (Kar et al., 2017). Luo et al. (2021) recruited 902 men for a cross-sectional study of 24 PFAS compounds in blood plasma. Five reproductive hormones: total testosterone (TT), E₂, FSH, LH, and insulin like factor 3 (INSL3), and SHBG were measured. They found a significant inverse relationship between PFAS mixture with E₂ and E₂/TT. Xie et al. (2021) studied a large cohort of males and females aged from 12 to 80 and found that PFAS exposure was associated with alterations in sex hormones (including TT, free testosterone, E₂, and SHBG) in a sex-, age-, and compound-specific manner. Multiple epidemiologic studies (Taylor et al., 2014; Barrett et al., 2015; Zhou Y. et al., 2017) associate PFAS exposure with effects indicating ovarian alterations (e.g., altered menstrual cycle, hormone alterations, early menopause). Petersen et al. (2018) examined PCBs and PFAS blood concentrations and associated effects on semen quality and reproductive hormones in Faroese men. Concentrations of total PCBs and PFOS were positively associated with sex hormone-binding globulin (SHBG) and luteinizing hormone (LH). While the PCB concentrations in the Faroese men were relatively high (compared to men around the world), the PFAS concentrations were relatively similar. Overall, the authors concluded that the positive association to LH for both PCBs and PFOS indicate a direct toxic effect on Leydig cells.

Petersen et al. (2020) qualitatively assessed 26 epidemiological studies for evidence of associations between PFAS exposures and male reproductive health. Several studies found some evidence that single PFAS compounds were associated with specific male reproductive parameters (e.g., semen quality, reproductive hormones, cryptorchidism, hypospadias, and testicular cancer). However, the overall conclusions were limited due to a lack of consistency between studies.

Data for short-chain PFAS effects on human reproductive parameters are scarce and these data are inconsistent (Nian et al., 2020; Brase et al., 2021). Indeed, the threat that these long-chain replacements pose on human populations, pathophysiology and health effects are not completely understood (Sunderland et al., 2019). Effects in humans are difficult to extrapolate from laboratory dose-response results. However, existing data indicates that concentrations in humans are lower than dose-response effects found in the laboratory. For example, So et al. (2007) found a maximum human PFAS concentration of 360 ng/L (breast milk), orders of magnitude below laboratory animal model exposures which induced toxicity.

Because of the likelihood of co-occurrence, short- and long-chain PFAS have been studied together. A three year study of 752 women in China found prenatal exposure to one short-chain and one long-chain PFAS (perfluorobutanesulfonate, PFBS and perfluoroheptanoic acid, PFHpA, respectively) was associated with alterations of fetal gonadotropins as well as free androgen levels (Nian et al., 2020). A recent review of endocrine disruption effects associated with both long- and short-chain PFAS exposure found that effects varied similarly according to gender and age of development; in some cases, short-chain effects were greater than long-chain PFAS (Mokra, 2021). While much more is needed pertaining to human reproductive effects, the implications of these studies warrant further epidemiological studies on short-chain PFAS.