Per and polyfluoroalkyl substances scientific literature review: water exposure, impact on human health, and implications for regulatory reform

Evidence acquisition

The objective of this review was to survey the approaches that identify the scope of the threat that PFASs present to humans through water contamination from a regulatory reform perspective. Sources for this review were found using the search terms “PFAS”, “PFASs”, “PFAS routes of exposure”, “PFAS water exposure”, “PFAS and drinking water”, and “PFASs and water” in several databases. These included the National Center for Biotechnology, ScienceDirect, PubMed, Elsevier, and Google Scholar. Themes and terminology used in scientific literature emerged from the search [6]. The review focused on research articles from 2017 to 2020 available online from 2015 to July 15, 2020. A complimentary search using the term “PFASs human health effects” was also conducted.

The focus of follow-up searches was on PFAS exposure and water, PFAS exposure and impacts on human health, and PFAS exposure in contaminated food. PFAS exposure was consistently linked to women’s health issues, and outcomes were often compounded. Therefore, health studies on the impact of women’s health were included. All searches were limited to articles for which full text was available in English and, for scientific articles, available in peer-reviewed journals.

A third search included general and specific terms for PFAS exposure from water other than drinking water or consumption. The search included multiple peer-reviewed journals: EBSCO, ProQuest, ScienceDirect, MEDLINE, PubMed, and others.

Lastly, a search for recent literature was conducted in Google Scholar and PubMed, focusing on geographic areas or populations that had produced few results from the previous searches. These searches included the terms “perfluoroalkyl substances OR PFOA OR PFOS AND” (for example) “Africa” OR specific country name for each country in the region for each of the following regions and populations: Central America, South America, Africa, the Middle East, and minority groups in the United States.

Selection of articles and studies and criteria of inclusion

Studies analyzing a wide range of matrices, including blood, urine, breast milk, and seminal plasma across different countries, were included to give a global picture of the health effects of PFAS exposure. Articles that covered the most recent developments in monitoring and standards served as focal points for the selection criteria, namely PFAS distribution in drinking water, regulatory frameworks, and legal implications. Countries of interest as the result of data that emerged included the United States, China, and countries in the European Union. Search results related to food exposure were restricted to examples directly linked to water contamination. Op-eds were excluded in the search to reduce bias and improve the reliability and trustworthiness of the data [7]. Several additional criteria were used to exclude articles. These include small sample sizes, sole authors, studies funded solely by companies currently facing multiple lawsuits as manufacturers (or former manufacturers) of PFASs, PFASs removal technologies, duplicates, PFASs measurement technologies, posters, presentations, no direct link to humans, and animal studies. Screening was performed by reviewing abstracts and the full text of the articles. The quality of the study was determined using conflict of interest statements, sample sizes, calculations for confounding variables, and statements provided in the studies by the researchers.

Potential biases may include publication bias and confirmation bias. Most of the research makes recommendations for further evaluation; many caution against wide use of PFAS to protect human health or recommend increased/stricter regulations, health advisories, or health guidelines.

Drinking water

Water is one of the direct pathways to contamination of per-and polyfluorinated substances. The closer a PFAS source is to a water supply, the greater the chance of pollution. PFASs in drinking water raise questions about potential health effects. The literature suggests that PFASs are present globally and are marked by differences in geography and production. Infrastructure is insufficient to limit its spread.

Domingo and Nadal conducted a literature review for the period from 2009 to 2019 focused on human exposure to PFASs through drinking water, including tap water and bottled water [55]. Overall, the authors concluded that the studies from around the world of PFAS exposure showed that PFASs in drinking water did not pose an imminent threat to human health or was of concern for regular consumers of tap water as measured by various (and widely diverging) health advisories, guidelines, and regulations.

Yet PFASs’ persistence in groundwater may impact the quality of drinking water for generations. Studies indicated that bottled water consumption is linked to lower PFAS levels in humans in areas with known or suspected PFAS-contaminated public drinking water systems [25], [26], [35], [55].

USA

According to Hu et al. low exposure to PFASs through drinking water can still lead to elevated exposures in humans. Elevated PFAS concentrations have been reported near industrial sites that produce PFASs or use them in numerous US regions. For example, PFOA concentrations were measured at 190 times the US EPA recommended lifetime health advisory (70 ng/L) in drinking water near a fluorochemical facility in Washington, West Virginia, that used PFOA during fluoropolymer production. Groundwater and surface water are reported to have 3–4 orders of magnitude above the US EPA health advisory level for drinking water for PFAS levels near many military fire training airports and civilian airports. Wastewater treatment plants are another source of PFASs because standard treatment does not remove PFASs from effluent, thus changing the concentrations of effluent relative to influent [41].

Hu et al. performed a spatial and statistical analysis on publicly available information about six PFASs in 36,149 drinking samples between January 2013 and December 2015. According to that study, 16.5 million Americans in 33 different states, three territories (American Samoa, Guam, and the Northern Mariana Islands), and the Salt River Pima-Maricopa Indian Community had PFASs present at or above the Maximum Recommended Levels by EPA in 194 of 4,864 public water sources. Drinking water for 13 states accounted for 75% of the frequency detection occurrences. PFASs were detected in drinking water from groundwater twice as much as that sourced from surface water. The study found a positive association in detecting all PFOS, PFOAs, PFHxS, and PFHpA, with statistical significance for PFHxS and PFOS in areas around military sites. They detected a small but significant increase in PFOA and PFOS in areas surrounding wastewater treatment plants. There was no association between proximity to the airport and elevated PFASs in drinking water near civilian airports certified to use aqueous firefighting foam (AFFF). The study noted that much data is missing for determining PFAS contamination in US drinking water. This is due to misclassification of airport AFFF eligibility versus actual use and poor information about PFAS release from smaller industrial facilities that often label this data as confidential business information [41].

Although a 2019 study found that exposure to consumer products with PFAS resulted in different body burdens of PFAS between African American women and non-Hispanic white only women, residing in an area with PFAS-contaminated drinking water was associated with higher levels of PFAS body burden, even though that group (n=6) was small, showing that PFAS contaminated drinking water poses a major public health threat [40]. A study of El Paso, Colorado, a community with known PFAS-contaminated water from AFFF from a nearby military site, supports contaminated public drinking water systems as the dominant pathway for exposure to PFAS in humans [35].

Sweden

Many studies were conducted to examine the effects of PFAS contamination in Ronneby, Sweden. Since the mid-1980s, one third of households received drinking water contaminated with PFAS, likely from AFFF from a nearby military airfield [32]. These studies found that PFHxS has a longer half-life compared to PFOA and PFOS and that drinking PFAS contaminated water resulted in higher serum concentrations and positive associations between PFAS exposure and serum lipids measuring cholesterol. There are potential associations between PFAS toxicity and downregulation of specific microRNAs, but they may have a high threshold of exposure. A study of residents of Ronneby also found evidence that linked raised risk of Crohn’s disease for certain exposure periods [30], [31], [32], [33]. Banzhaf et al. outlined 11 challenges to Sweden in addressing PFAS water contamination, which may impact drinking water for more than 3.6 million individuals. They identified a communication shortage among Swedish regulating authorities and researchers to establish clear guidelines for monitoring PFAS contamination [56].

Italy

Turning to Italy, a three-year monitoring campaign from 2010 to 2013 in northern Italy by Castiglioni, et al. indicated that there was widespread groundwater contamination by PFAA, but that drinking water treatments for Milan removed these contaminants and drinking water to consumers was far below proposed thresholds for PFAS [24]. In contrast, the highest levels of PFOA were detected in the River Po. The study identified wastewater treatment plants as a source of PFAS pollution, particularly when the plant received industrial waste [24].

Mastrantonio et al. conducted an ecological mortality study in the Veneto Region of Italy to assess the differences among general populations residing in areas with PFAS contaminated groundwater versus unaffected areas [27]. Areal selection criteria were based on monitoring reports and campaign results for PFOS, PFOA, and other PFAS conducted between 2013 and 2015 by environmental, public health, and government agencies in the region. Overall, 24 contaminated and 56 uncontaminated municipalities were represented in the study. Specifically, the research team delineated the following four factors that could influence mortality rates: smoking habits, socioeconomic status, sex, and age. Using epidemiological data from the National Agency for New Technologies, Energy and Environment (ENEA), three calculations were applied in the study areas for each cause of death recorded between 1980–2002 and 2003–2013: standardized mortality rates, sex-specific number, and rate ratios (RR) [27].

The researchers found that elevated mortality rates for some causes of death, potentially associated with PFAS exposure, occurred in contaminated municipalities in contrast with uncontaminated ones with similar socioeconomic status and smoking habits. In contaminated municipalities, statistically significant rate ratios for both males and females were found for the following conditions: general mortality (1.19 in males; 1.21 in females), diabetes (1.21 in males; 1.48 in females), cerebrovascular diseases (1.34 in males; 1.29 in females), Alzheimer’s disease (1.33 in males; 1.35 in females), and myocardial infarction (1.22 in males; 1.24 in females). Although RR values were greater than 1 for bladder cancer and leukemia for both sexes, they were not deemed statistically significant. For females, RR values were significantly greater than 1 for Parkinson’s disease (1.35), kidney cancer (1.32), and breast cancer (1.11). For males, no statistically significant risk was found for pancreatic cancer, testicular cancer, and Parkinson’s disease [27].

Moreover, no increased risks were found for several serious diseases, such as liver cancer, non-Hodgkin’s lymphoma, and prostate cancer. However, the researchers did observe a significantly lower risk of liver cancer when males and females were considered together. Lastly, myocardial infarction risks were 13–45% higher in communities living near a PFOA chemical plant. The research team concluded that the results “warrant further individual level analytic studies to delineate causal associations” [27].

After alarming levels of PFASs were found in the Veneto Region’s drinking water in 2013, Italy’s National Health Institute (ISS) decreed the following industry performance limits in 2014: PFOS: ≤30 ng/L; PFOA: ≤500 ng/L; other PFAS: ≤500 ng/L [25]. Following this precaution, Ingelido and her team at ISS compared exposed and non-exposed groups from the general population residing in the Veneto region in an area near a plant that produced PFAS for decades and where drinking water contamination with PFAS was discovered. The human biomonitoring study was conducted by measuring PFAS concentrations, particularly PFOA, in human sera by taking blood samples from residents in the region and comparing those to concentrations in neighboring populations at background PFAS exposure. In addition, the population was genotyped for the OATP1A2*3 allelic variant. The PFAS concentrations measured in human sera were correlated with drinking water consumption by aqueduct and private well. Other variables from a questionnaire on lifestyle habits were also considered. Ingelido et al. found that the difference in PFAS concentrations between exposed and not exposed groups was significantly higher for 9 of the 12 PFAS substances analyzed [25]. This finding confirmed water contamination had resulted in a high PFAS exposure of the residents over time. Ingelido et al. found that the main factors influencing PFAS serum levels was residence in the area and the related extent of drinking water contamination [25]. Other factors that played a role included sex (male PFAS levels higher than female PFAS levels), water consumption (higher consumption of tap water correlated with higher levels of serum PFAS concentrations), years of residence (in areas with contaminated water), and raising own livestock. The study found no relationship between the genetic trait for renal transport (OATP1A2*3) and PFAS levels. Age did not seem to strongly impact levels of PFASs [25].

In a follow-up biomonitoring study, Ingelido et al. compared a subgroup of 122 farmers living in rural areas in the Veneto region, where public drinking water contamination was discovered in 2013–2017. The farmers produced and consumed their own livestock and vegetables. They frequently used well water. Ingelido et al. compared these farmers exposed to PFAS drinking water contamination and people not exposed to drinking water contamination in the Veneto region [26]. Ingelido et al. compared PFAS levels in human sera; divided the area into subareas depending on proximity to PFAS groundwater plumes; segmented the population based on gender, age, and whether the farmers grew and raised their own food; performed a genetic analysis of an allele, and considered information from a questionnaire. Ingelido et al. found that the highest levels of PFAS were in exposed farmers in the rural contaminated areas. They concluded that the main factors influencing PFAS levels in farmers were “residence area and the related extent of drinking water contamination, gender, years of residence in the municipalities, well water consumption and consumption of own produced food” [26].

Ingelido et al. found higher concentrations of PFAS in males than females for PFOA, PFHxS (5 times as much) and PFOS (3 times as much) when comparing median levels [26]. When comparing exposed farmers for PFOA, that subgroup had three times as much PFOA in human sera compared to exposed groups in the region and 25 times as much to the non-exposed group. Their research seemed to support that individuals consuming their own livestock were linked to higher PFAS levels in human sera. A limit of the study was a self-reported data questionnaire for amounts of food eaten. Nonetheless, they also concluded that the primary source of PFAS exposure to people in the Veneto region was consuming contaminated public drinking water. As an example of the main factors that influenced PFAS levels in humans, the researchers mentioned that the highest concentration of PFOA was found in a 44 year old man who lived in a contaminated area (ULSS 5) who drank contaminated tap water and grew his own fruits, vegetables, and animals raised with contaminated water [26].

China

In China, Yan et al. found a strong link for PFASs between groundwater near landfills and tap water supplies [14]. Moreover, PFASs were found to be almost ubiquitous in drinking water across China, with levels ranging from 4.49 to 179.93 ng/L. Other sources contributed to these values, including proximity to fluorochemical manufacture sites and water sources in or near metropolitan areas [14].

Duan et al. investigated the general distribution of PFASs, including ultrashort and short-chain PFASs, in Tianjin, China, finding that PFOS alternatives were widely distributed (> 90% of samples). Long-chain and ultrashort PFCAs (∑PFCAsC2–C3), accounted for 70 and 28.4% of PFCAs respectively in the human body burden as measured in human sera. They reported a significant positive association between serum PFAS concentrations and HbA1c, a glycemic biomarker, in people 55 years or older [9]. Several studies found adverse impacts on prenatal, neonatal, and natal health outcomes and PFAS exposure, including replacement PFASs [8], [12], [17]. Huo et al. did not find a significant and consistent pattern of association with PFAS exposure and gestational hypertension, preeclampsia, or hypertensive disorders of pregnancy [10]. Associations between PFAS body burden and visual impairment diseases in China and increased risk of breast cancer, particularly among younger Taiwanese women were also found [15], [18].

Faroe Islands, Norway, Finland, Denmark

The studies covered in this review conducted in Norway and the Faroe Islands lent support to evidence that PFAS may have immunosuppressive effects [21], [22], [29]. Studies conducted in Denmark found associations between PFAS exposure during childhood two outcomes: adiposity in adolescence and young adulthood, and effects on puberty onset [19], [20]. A study from Finland noted that body burdens are a better indicator of PFAS than serum concentration in children as it accounts for dilution, noting an increasing temporal trend in PFHxS and PFNA during early childhood [23].

Various countries, Africa, Central America, South America

Some studies tested bottled and tap water samples from around the world for PFASs, which helped to allow comparisons. Overall, bottled water bought at markets around the world had lower levels of PFASs compared to tap water [54]. PFAS concentrations in tap water in Burkina Faso and the Ivory Coast were also found to be comparable to PFAS concentrations in Canada [54, 57]. A critical review of environmental levels and human body burdens noted that the two studies of drinking water in Africa encompass three countries and indicate a broader problem in Africa of potential PFAS contamination in tap water from nearby industries and wastewater treatment plants [57]. By contrast, Ahrens et al. concluded that the levels of PFOS in fish from Tana Lake, Ethiopia did not present a risk to the fish-eating population [53].

There are few studies of PFAS contamination from Central and South America. However, a recent study assessing whole blood in Brazilian pregnant women found a significant positive association between PFOS and PFOA and fetal growth restriction [51]. A study from Denmark and Faroe Islands found evidence suggesting links between PFAS levels and increased odds of asthma in children at ages 5 and 13 [52].

Recreational exposure to PFAS

The Environmental Work Group (EWG) and the Social Science Environmental Health Institute of Northeastern University estimate that 19 million Americans are affected by contaminated drinking water [58]. However, concerns have arisen about additional pathways, including recreational and dietary exposure. Concerns over exposure to emerging contaminants have been well documented and publicized from recent accounts of ground and municipal drinking water contamination [58]. However, the evidence does not support a severe risk of dermal exposure to PFASs from swimming or other recreational water-related activities, beyond the risk posed from accidental ingestion of contaminated water or foam. This contrasts with the dangers associated with dermal and inhalation exposure seen in occupations such as carpet installation, firefighting, and other sources. As a result, the presence of PFASs in both the natural and industrialized food chain adds a significant risk to the overall exposure of PFASs to affected residents, including exposure from mother to child through breast milk [59].

Dietary & food source exposure to PFAS

Fishing may be one pathway of PFAS exposure. Different rates of elimination between humans and fish will require new methods of assessing evaluations of weight-of-evidence to comprehend bioaccumulation mechanisms across different hierarchical levels in ecological pyramids for PFASs. To add to this, the accumulation of PFAS in the environment will lead to increased external exposures [3].

Studies show that low-income households consumed more self-caught fish than those with higher income and education [60]. The results of a US survey indicated that recreational or self-caught fish consumption rates vary regionally and are not well understood, making it difficult to support national-scale assessments. Low-income respondents were more likely to harvest fish as a food staple and reported the highest consumption rates. In contrast, the higher income and education respondents reported higher incidences of self-caught fish but overall lower consumption rates. There were also regional differences. Respondents from the East-South Central and New England regions reported the lowest consumption rates from self-caught fish (12 to 16 g/day). By contrast, Mountain, Pacific, and Mid-Atlantic reported the highest consumption rates from 44 to 59 g/day. Respondent-specific consumption rates and national-level data on fish tissue concentrations of PCBs, MeHg, and PFOS suggest that 10–58% of respondents reporting self-caught fish consumption are exposed to concentrations of PFOS at levels that exceed threshold levels for health effects, estimated to represent 2.3 to 19 million people [60].

Fish consumption may also differ across cultural cuisines with differing levels of exposure to PFASs. PFASs are found in freshwater arctic wildlife, which are traditional food sources for indigenous peoples of the region [39]. In addition, a recent longitudinal study of PFAS body burden in US midlife women investigating differences on the basis of race/ethnicity, parity (parous/nulliparous), and menstrual status noted that Chinese women had higher levels of PFNA as a baseline and temporally [40]. This may relate to diet (fish consumption).

Short-chain PFASs

According to Brendel et al. studies have linked long-chain PFASs to either toxic for reproduction or persistent, toxic, and bio accumulative properties or both [61]. Long-chain PFASs are subject to regulation under REACH (EC No. 1907/2006). By contrast, short-chain PFASs do not fall within the regulatory ambit of REACH. As a result, short-chain PFASs are widely used as an alternative to long-chain PFASs. This is in part because short-chain PFASs are not regulated as persistent organic polluters (POP) under measures such as the Stockholm Convention or the European Chemicals Regulation (REACH EC No. 1907/2006) [61].

Short-chain PFASs are persistent to the same extent as long-chain PFAAs and widely distributed in the environment. There is not a lot of information about short-chain PFASs as they are not regulated. Compared to long-chain PFAAs, short-chain PFAAs have a greater potential for long-range mobility. Brendel et al. concluded that drinking water resources were extremely sensitive to short-chain PFAAs contamination. Because of low adsorption potential, short-chain PFAAs do not bind with other chemicals, so they remain dissolved in water, including drinking water resources [61]. Short-chain PFAAs in contaminated drinking water resources lead to exposure in humans that is poorly reversible. This lack of removal leads to continuous exposure through the drinking water, which is independent of short-chain PFAAs’ elimination half-lives (Brendel et al. 2018). Because of the increased use of short-chain PFASs, regulating them is imperative under REACH in the European Union [61].

Research on these traditional PFAS alternatives is limited, even though there are some indications that they produce similar adverse health effects as their predecessors and are even more fluid within the environment. Some studies suggest even greater dangers of toxicity [62], [63]. Temkin et al. examined PFASs, including short PFASs, from a toxicological standpoint using the Key Carcinogens framework. They also noted the lack of studies on PFOS and PFOA alternatives, such as short-chain PFASs [64].

Long-chain PFASs

Post, Gleason, and Cooper conducted a study in 2017 detailing the toxicological effects of exposure to long-chain PFAAs. Breastfed babies are at a higher risk than older individuals, even from low contamination levels to which the lactating mother is exposed. This is because long-chain PFAAs are passed many times over through the mother’s breastmilk to the infant [65]. Thus, infants are at a greater risk of developing adverse developmental health effects in their earlier, more sensitive stages of life.

Adults are not without health effects from long-chain PFASs exposure, however. Zeeshan et al. investigated the relationship between PFAS in human blood and eye diseases, including visual impairment in the Chinese population, concluding that risk of the vitreous disorder increases with long-chain PFAS exposure [15].

General populations

Australian National University’s National Centre for Epidemiology and Population Health, Research School of Population Health, reviewed 221 pieces of published literature to determine if PFAS exposure influenced health outcomes. They conducted their review in 2017, and at that time, their investigation assessed PFASs association of health outcomes across 12 categories. These included skeletal, immunological, body weight, cardiovascular health, cancer, diabetes, neurological, thyroid, metabolic, reproductive, maternal health, infant health, neonatal health, and respiratory health. The team used a multi-domain risk of bias tool to evaluate their sources. Their evaluation found adequate proof that PFAS exposure increased blood cholesterol by a low but significant amount. They found additional but limited proof of PFAS links to kidney disease, kidney cancer, testicular cancer, high uric acid concentration in the blood, and immunity for rubella and diphtheria following vaccination [64].

In 2020, because PFAS has been identified as a probable human carcinogen, a novel, state-of-science review by Temkin et al. applied key characteristics analysis into risk assessments. This approach is being increasingly used in the USA and international government agencies, to data for several PFAS (26 chemicals) for each characteristic to evaluate the strength of evidence as a carcinogen [64]. Temkin et al. applied the Key Characteristics of Carcinogens framework. They identified differences in the strength of evidence between long-chain and short-chain PFAS. For example, Temkin et al. found strong evidence of oxidative stress and suppressed immune response for PFOA and several long-chain PFASs, but only suggestive evidence for one short-chain PFAS for those health effects. Well-studied PFAS such as PFOA and PFOS were found to have up to five key characteristics of carcinogens. Temkin et al. noted that although over 600 different PFAS are known to be inactive use over the last decade, for most of those compounds, no toxicological information is available [64].

A study by Duan et al. found novel PFASs PFAES and TFA widely detected in high concentrations in an adult population in Tianjin, China [9]. Jin et al. found that an isomer of PFAES concentrations increases with age in a Chinese population in Anji, China [11]. Among Chinese adults, Zeeshan et al. observed that blood levels of PFASs were significantly higher in a group with eye disease when compared to a group without such disease. A recent study found evidence suggesting an association between PFAS and uric acid varied according to adult kidney function and isomer [15].

Prenatal, maternal, infant

Maternal, prenatal, and infant exposure to PFAS will be addressed together as they are inextricably interrelated pregnancy, delivery, and lactation are known excretion routes for women, and serum concentrations dropped with increasing deliveries [28].

Although PFOA and PFAS were not found to be associated with hypertensive disorders during pregnancy, PFHxS was positively associated with preeclampsia in a study with 1,739 participants from the Maternal-Infant Research on Chemicals (MIREC), a longitudinal cohort of pregnant Canadian women [34]. The Borghese et al. 2020 study may have found the strongest effect of any PFAS observed to date; they observed that the risks of developing preeclampsia were 3-fold higher among pregnant women in the highest (relative to lowest) tertile of PFHxS. This finding may have been a result of using a definition of preeclampsia as defined by guidelines by the Society of Obstetricians and Gynecologists of Canada instead of the American College of Obstetricians and Gynecologists.

Because studies have shown that alterations in thyroid function in the mother during pregnancy are associated with several adverse fetal outcomes such as preterm delivery, insufficient fetal growth, and neurodevelopment deficits, Preston et al. investigated the recent epidemiological studies’ suggestion that PFASs may alter maternal and neonatal thyroid function [45]. Specifically, Preston et al. sought to assess the effect on maternal and neonatal thyroid function from exposure to PFASs individually and in mixtures in a prospective cohort in Boston by using two statistical methods: weight quantile sum regression (WQS) and Bayesian kernel machine regression (BKMR). Using both statistical methods, Preston et al. found that the PFAS mixture was inversely associated with maternal FT4I and neonatal T4 levels, primarily in male infants. The researchers noted the BKMR results identified other associations and that BKMR and WQS regression analysis were relatively consistent. The use of BKMR permitted analysis of potentially non-linear exposure-response functions and interactions among PFAS. This is one of the first epidemiologic studies, if not the first, to examine PFAS mixture effects on maternal and neonatal thyroid function [45].

Although the emerging literature on prenatal PFAS exposure and neurodevelopment is still in its nascence and so inconsistent, a study from 2020 show associations between prenatal exposure to PFAS and neurodevelopment [48]. Studies examining prenatal PFAS exposure and IQ and other childhood cognitive outcomes have had inconsistent results, some adverse on IQ while others were positive. These developmental effects and delays have been found to last years into childhood. Spratlen et al. found evidence suggesting sex-specific associations between prenatal exposure to PFASs and childhood neurodevelopment; specifically, higher prenatal PFAS exposure may be associated with higher scores for some cognition for females but not for males. Spratlen, et al., noted that these findings should be interpreted with caution due to limited data and statistical tests and citing different results from other studies [48].

PFOS and PFOA can cross the placental barrier between mother and child and alter the size, birth weight, Intrauterine Growth (IUGR), and gestational age [51]. Preterm birth, or Intrauterine Growth Restriction (IUGR), has been associated with increased risk for several diseases later in life, including cardiovascular disease, obesity, and endocrine disorders. A study in Brazil examined the potential links between PFASs concentrations in the whole blood samples of 252 pregnant mothers and preeclampsia, low birth weight, preterm birth, and IUGR [51]. The study in Brazil found no significant association between increasing PFOS and PFOA and preterm birth, preeclampsia, and low birth weight. The bivariate analysis, however, showed a positive association between concentrations of PFOA and PFOS and IUGR. Souza et al. 2020 also found a significant association between socioeconomic status and PFOS/PFOA levels; more affluent women had higher concentrations of PFASs. Whole blood concentrations of PFOS and PFOA were also significantly associated with fetal growth restriction in pregnant Brazilian women [51].

The maternal serum concentration of an alternative to PFOS used in China (Cl-PFESA) was associated with lower birth weight and higher risk for preterm birth [8]. A stronger association between PFAS and gestational age among female infants was found, the reason for the sex-difference was unknown. Kashino et al. also found that prenatal concentrations of PFAS were linked with adverse birth outcomes, including birth weight [17].

In short, there is support in the literature for the notion that humans excrete PFASs through breastfeeding [46]. A 1-month-old formula-fed baby consumes 137 mL of drinking water per kg in a day, while, by contrast, an 11-year child consumes 13 mL per kg per day [68]. As a result, infants show elevated serum PFAA levels because they are exposed to more contaminants [65].

PFAS exposure and children

According to a 2017 literature review of 64 studies that address links between PFAS exposure and child health outcomes done by Rappazzo et al. six outcomes emerged, namely immunity/infection/asthma, neurodevelopmental/attention, thyroid, cardio-metabolic, renal, and puberty onset [2].

Regarding asthma, a 2020 study by Kvalem et al. had two designs: cross-sectional and longitudinal and investigated PFAS exposure and asthma. They noted that no firm conclusions could be drawn about PFAS exposure and asthma and asthma-related health outcomes, because, although cross-sectional studies (including theirs) do link PFASs and asthma health outcomes, longitudinal studies of PFASs’ relationship to asthma (including theirs) do not find a link (or a clear gender pattern) [29]. Nonetheless, Kvalem et al. did find a pattern of a positive association between allergic sensitization and PFASs in all participants, regardless of study design (cross-sectional or longitudinal). Despite that finding, they concluded that no conclusion could be drawn between PFASs and rhinitis or allergic sensitization as there are inconsistent findings and too few studies. The longitudinal study found an overall trend between increased PFAS levels in human serum and lower respiratory tract infection between the ages 10 and 16 for all participants, which varied based on gender. Kvalem et al. found that the increased risk of lower respiratory tract infection in the longitudinal study at different exposure periods suggested a PFAS exposure may induce immunosuppression both after prenatal and childhood exposure [29].

Several studies add to the growing evidence that children, especially when young, are particularly vulnerable to PFAS exposure, particularly for immune responses [21], [22]. A recent NHANES study collected information about PFAS concentration in children’s serum, finding that PFAS concentrations were generally higher in children whose samples were collected in the mid to late 2000s, after regulated PFASs such as PFOS and PFOA were in decline. Interestingly, the researchers found that Mexican Americans had lower PFOS and PFOA concentrations than non-Hispanic persons [50].

PFAS exposure and adolescents

A study found that exposure to PFOS during childhood was associated with increased adiposity in adolescence and young adulthood [19]. Ernst et al. observed that PFAS concentration in maternal plasma was associated with lower age at puberty marked onset for certain PFASs but later puberty onset for other PFASs [20].

Women

As ovarian function can indicate overall health in midlife women, a comprehensive literature review was conducted to assess available information about PFAS exposure and ovarian function [68]. The findings from the studies suggest that PFAS exposure can adversely affect ovarian function, including essential functions such as folliculogenesis and steroidogenesis, to which girls compared to adults may be more vulnerable. Despite the growing body of literature, the review concluded that the inconsistencies of previous findings from studies and lack of longitudinal evidence prevented making causal inferences at this time [68].

Ding et al. performed a longitudinal study that temporally observed PFAS trends in human serum concentrations comparing race/ethnicity, parity (parous/nulliparous), and menstruation in a national cohort across the USA among White, African American, and Chinese women [40]. The comparisons by race/ethnicity were novel and showed differences in body burden and rate of an increase or decrease in PFASs over time. Also, to our knowledge, this was the first longitudinal study of midlife women to compare PFAS concentrations between parous and nulliparous women; it found that parity was a significant determinant of serum PFAS concentrations. Parous women had lower PFAS concentrations. The study also strengthened the evidence that menstruation status influences serum PFAS concentration; menstruating women had lower PFAS concentrations. In sum, the results depict a decline in legacy PFAS (PFOA, PFOS) and their branched isomers and precursors among women living in the USA during 1999 and 2011. This is consistent with reduced environmental burden [40].

In a case-control study, Tsai et al. examined the relationship between PFAS exposure and risk of breast cancer, stratified by age group (>50, ≤50) [18]. Tsai et al. noted that the few studies between breast cancer and PFAS were in the USA and Europe, and stated that more research was needed in “ethnic populations” [16]. The results showed that younger women had a higher risk of breast cancer based on PFOS exposure compared to older women (>50), which differed from other studies. They also found a higher risk of women 50 and younger of ER-positive tumors compared to women over 50 due to PFHxS and PFOS exposure. In contrast, young women had a lower risk of ER-negative tumors compared to older women based on PFNA and PFDA exposure. They noted that because the recruitment of participants was cross-sectional, causal inferences could not be made [18].

Regulatory frameworks, information gaps, risk assessment

Because many standards are voluntary, there is still a need for enforceable regulations on PFAS limits in drinking water. Although some governments and regions are moving towards regulating PFAS with enforceable limits, a constellation of different health advisories, guidelines, and provincial regulations have emerged Table 5.

The US regulatory framework serves as a microcosm of the myriad of considerations in assessing risk and implementing guidelines and regulations. In assessing the disparity regarding health advisories, health-protective guidelines, and standards to limit exposure to PFASs via drinking water, three underlying influences are noted: scientific uncertainty; varying degrees of risk assessment decisions; and an array of economic, social and political factors [5]. In numbers, the range of levels considered safe by different jurisdictions spans from 12 ng/L to 600 ng/L [5]. The final limit on PFAS in drinking water reflects a series of compromises in creating regulation among many stakeholders who consider many factors.

Cousins et al. proposed several approaches from a sound, scientific perspective to set regulatory standards. They discussed how to handle current knowledge gaps, cost considerations, current implementation of the different approaches used today, and the danger to human health and the environment from PFASs regulations. Different jurisdictions must decide which approach is suitable [72]. However, Cousins et al. point out, a broader approach that considers scientific uncertainty and knowledge gaps will require less expertise to apply. They will likely incur less expensive to implement while protecting human health and the environment under current circumstances. Based on the research for this review, we believe that an insistence on a chemical-by-chemical approach to regulating PFASs appears to have a sunset in regulatory approaches for two reasons: 1) public awareness about PFASs’ pervasiveness and (at least some) adverse health effects; 2) the fact that PFASs’ creation (estimated 4700+) has outpaced regulation and scientific research (2 consistently PFOA/PFOS) for decades. The USA’s recent direct legislation of PFAS requiring notification for releases in the USA and the EU’s recent actions regarding PFASs and REACH in 2020 serve as support for our belief. Furthermore, it seems that broader regulation of PFASs is a question of how as opposed to whether and when.

While having one global standard for PFAS contamination in drinking water may forever be impossible under current structures, at the very least, clear national standards should be set in every country, if not regional organizations. Otherwise, like judicial forum shopping in which a forum participant seeks the jurisdiction with the rules most favorable to the participant, industry stakeholders may continue to engage in regulatory forum shopping to set up PFAS manufacturing facilities, thus thwarting community efforts to manage PFAS contamination risks to human health. Authors reporting a recent community-led health survey of residents in Merrimack, New Hampshire, a place where a local chemical plant led to PFAS-contaminated public drinking water and private wells, the authors seem to have outlined such a regulatory-forum shopping approach by an industry stakeholder [73]. Although national and local jurisdictions have borders, water sources often cross those borders. As a result, clear standards covering larger geographic areas are necessary. This approach would benefit communities as well as industry stakeholders, who would then have clear rules to follow instead of risking litigation and its resulting uncertain legal and financial liabilities. To take one historical example, the C8 panel, and the funding for its numerous longitudinal studies resulted from the settlement of a lawsuit. A strong public awareness campaign about PFAS to targeted groups, such as people who are likely to eat harvested fish, is also needed in the public health approach.

From a scientific perspective, studies that have considered drinking water limits also provide a range of possibilities. To protect against potential immunosuppressive effects in children as evidenced by antibody responses to vaccines, one benchmark dose study recommended 1 ng/L limit for PFOA [21]. Another study noted the European Food Safety Authority Daily Intake limit and applied the World Health Organization conversion rules to state a limit of 3000 ng/L for PFOA and 300 ng/L PFOS [24]. In 2019, a report by scientists assessing limits for Michigan’s public water drinking systems urged that scientific evidence supports a need for a maximum contaminant level in drinking water of 0 for all PFASs [63].

Inconsistent and insufficient monitoring

While the lack of standards presents a problem for global comparison and consensus, it also presents a gap of evidence with more regional and community-based implications. The variability in standards and observed values is particularly problematic when issuing guidance and advisories for PFAS-related health issues across multiple geographies. Moreover, testing protocols within each territory may not reflect the complete scope of PFAS contamination of community water sources. For example, Michigan has well-documented efforts to identify PFAS contamination in over 100 public water systems; yet their testing protocol excludes small private wells that are not designed for public use [63]. Thus, the extent of PFAS exposure to Michigan residents is still unknown.

Recent biomonitoring studies by the Centers for Disease Control (CDC) and the United States National Health and Nutrition Survey (NHANES) found it paramount to include biomonitoring for multiple PFASs beyond PFOS and PFOA to include PFDoA, PFNA, MeFOSAA, PFUA, PFDeA, PFHpA, PFBS, FOSA, EtFOSAA, and PFHxS [63].

Although Cl-PFESA is a widely used PFOS alternative in China, there is no human data available to assess the reproductive toxicity of this PFAS [8]. As Temkin, et al. noted, for the majority of 600 PFAS in wide use for the past decade, there is no toxicological data [64].

Rapazzo et al. note two considerations for future studies of PFASs and childhood health outcomes: longitudinal studies and variations in outcome measurement [2]. These authors of the Australian report observed a significant potential risk (potential) of bias based on factors such as study design and researcher motivation [66].

Observations

The design of the studies, the reporting measurements, statistical models, and other factors makes comparing results complex. Many of the studies relied on interviews of participants to gather information about lifestyle habits to identify covariates; some studies noted that participants might have recollection bias while others noted the possibility of underreporting in examining differences between self-reports of lifestyle behaviors Moreover, self-reporting may be faulty. Thus, many studies that identified covariates through interviews without confirmation with other records may have a bias.

As Ding, et al. noted, the studies also vary in how they report human body burdens of PFAS (e.g., ng/mL in serum versus odds ratio versus risk ratio versus ng/kg bw/day) which makes it difficult to compare the results across the studies [68].

Several studies also noted observations about the sample and the ability to generalize the results. Recent studies noted a lack of ethnic/racial diversity in study cohorts that may limit the generalizability of the findings to general populations that are not as homogeneous as the study cohort [45]. Tsai et al. noted that PFAS burden differs across ethnic groups in which PFAS are found in higher concentrations, it may be helpful to ensure a diverse cohort in addition to a more robust study of different PFAS for future studies of PFAS exposure and links to maternal, prenatal, and neonatal health outcomes [18]. Future studies conducted in the USA should include persons of Hispanic/Latinx ethnicities and Asian Americans [37], [40].

Although all studies used accepted statistical methods and approaches, the use of multiple statistical tests may have resulted in false positives, as several studies noted. Moreover, the wide array of statistical models both for analyses and adjustments make comparing results across studies difficult because, as one study showed Preston et al. 2020, results of associations changed depending on which statistical approach was used [45]. Although this review focused on the time period between 2015 and 2020, older studies were also reviewed, some of which were included. It was observed that older studies did not summarize in as much detail as studies from the past five years the statistical method calculation process.

The potential for reverse causality and the necessary need for caution in interpreting results recurred in many biomonitoring studies, particularly when interviews without access to medical records were performed at the time blood samples were taken. Moreover, study design itself may also affect which and whether associations are found (e.g., asthma and cross-sectional studies) [29].