Highlights

• •Exposure to PFAS through drinking water is a global human-health concern.
• •Paucity of information on PFAS exposure in unregulated private-well tapwater (TW)
• •Assessment of PFAS in > 700 private and public-supply TW in the United States (US)
• •Models suggest at least one PFAS detected in 45% of US drinking water samples.
• •Benchmarks and US proposed PFAS regulations exceeded in private and public-supply TW.

Abstract

Drinking-water quality is a rising concern in the United States (US), emphasizing the need to broadly assess exposures and potential health effects at the point-of-use. Drinking-water exposures to per- and poly-fluoroalkyl substances (PFAS) are a national concern, however, there is limited information on PFAS in residential tapwater at the point-of-use, especially from private-wells. We conducted a national reconnaissance to compare human PFAS exposures in unregulated private-well and regulated public-supply tapwater. Tapwater from 716 locations (269 private-wells; 447 public supply) across the US was collected during 2016–2021 including three locations where temporal sampling was conducted. Concentrations of PFAS were assessed by three laboratories and compared with land-use and potential-source metrics to explore drivers of contamination. The number of individual PFAS observed ranged from 1 to 9 (median: 2) with corresponding cumulative concentrations (sum of detected PFAS) ranging from 0.348 to 346 ng/L. Seventeen PFAS were observed at least once with PFBS, PFHxS and PFOA observed most frequently in approximately 15% of the samples. Across the US, PFAS profiles and estimated median cumulative concentrations were similar among private wells and public-supply tapwater. We estimate that at least one PFAS could be detected in about 45% of US drinking-water samples. These detection probabilities varied spatially with limited temporal variation in concentrations/numbers of PFAS detected. Benchmark screening approaches indicated potential human exposure risk was dominated by PFOA and PFOS, when detected. Potential source and land-use information was related to cumulative PFAS concentrations, and the number of PFAS detected; however, corresponding relations with specific PFAS were limited likely due to low detection frequencies and higher detection limits. Information generated supports the need for further assessments of cumulative health risks of PFAS as a class and in combination with other co-occurring contaminants, particularly in unmonitored private-wells where information is limited or not available.

Results and discussion

3.1. Spatial and temporal assessments of PFAS exposure in tapwater

In the US and globally, limited information is available on PFAS in point-of-use tapwater, with most drinking-water studies focused on samples from source waters (McMahon et al., 2022, Sims et al., 2022) or pre-distribution samples from community water supplies (Andrews and Naidenko, 2020, Hu et al., 2016, Kurwadkar et al., 2022, Li et al., 2022, McMahon et al., 2022, Neuwald et al., 2022, Post et al., 2013), largely omitting distribution system factors (e.g., plumbing material with PFAS or sorption/degradation in the supply network; (Mohammadi et al., 2022) and a notable paucity of data available for private-wells across the US. To address this gap, we utilized targeted analysis of up to 44 PFAS in point-of-use tapwater from 269 private-wells and 447 public-supply collected 2016–2021 as fractional indicators of the presumptive 8000 + PFAS contaminant (U.S. Environmental Protection Agency, 2022a) (Fig. 1, Figure S1, Table S1). Consistent with other studies and large datasets focused on public–supply tapwater (Hu et al., 2016, Li et al., 2022, McMahon et al., 2022, Post et al., 2013), at least one PFAS was observed in 30% (237 of the 716) of the tapwater samples collected throughout the US (Fig. 1, Figure S2). Based on data from the UCMR3, about 4% of US drinking-water treatment plants tested had detectable PFAS but the breadth of contamination was likely missed due to high detection limits (10–90 ng/L depending on individual PFAS) and a limited number of PFAS analyzed (Hu et al., 2016). More recently, assuming lower detection limits, Andrews and Naidenko (2020) estimated that approximately half the US population likely receive water with PFOA/PFOS concentrations<1 ng/L, but this information has yet to be validated fully with field data particularly at the point-of-use. In our study, seventeen PFAS were detected at least once and, apart from perfluoropropane sulfonic acid (PFPrS), all detected PFAS were analyzed in > 600 samples (Table S2). Reporting limits for the PFAS observed in the current study varied by laboratory and ranged from 0.1 to 20 ng/L (Table S3). The most frequently detected PFAS analyzed by all laboratories included PFBS (16%), PFHxS (15%), and PFOA (14%), similar to results reported by others for drinking-water resources (surface water intakes or groundwater wells) or community water supplies (Andrews and Naidenko, 2020, Boone et al., 2019, McMahon et al., 2022, Post et al., 2013, Teymoorian et al., 2023). The number of individual PFAS observed ranged from one to nine (median of two) with detected concentrations ranging from 0.025 to 319 ng/L (median: 2.88 ng/L) and corresponding cumulative PFAS concentrations (sum of 16 detected PFAS) ranging from 0.348 to 346 ng/L (median: 7.00 ng/L; Table S10). As expected, sites classified as ‘low’ impact had the lowest prevalence of PFAS compared to sites near known PFAS sources, whereas PFAS varied widely among sites classified as ‘medium.’.

At least one PFAS was detected in 20% of private-well (55/269) and 40% of the public-supply (182/447) samples collected throughout the US. A similar pattern was reported in groundwater from the eastern US, in which 60% of the public-supply wells and 20% of monitoring wells contained at least one PFAS (McMahon et al., 2022). Median cumulative PFAS concentrations (estimated considering detection limits and including study as a random effect) were comparable between public-supply (median = 7.1 ng/L [95% CRI = 2.3, 17.1]) and private–well point-of-use tapwater (median = 8.2 ng/L [95% CRI = 2.6, 20.5]; Fig. 2). Similarly, considering only PFAS with > 5% detections, we observed no differences in estimated median concentrations of individual PFAS or in the number of detected individual PFAS between public-supply and private-well point-of-use tapwater samples (Fig. 2, Tables S7-8). We chose 5% as a cutoff because this provided enough observations above the detection limit to allow for meaningful inferences of estimated parameters. Based on model predictions (Figure S3), the probability of not detecting PFAS above our detection limits ranged from approximately 25% in urban centers (e.g., Chicago) or areas with a known history of PFAS contamination (e.g., Cape Cod (Bradley et al., 2021a)) to > 75% in rural areas (e.g., Northern Plains (Bradley et al., 2022)). Figure S3 illustrates the spatial (among-study) variability that exists in PFAS occurrence (and concentration, not shown) quantified by the study random effect included in all models to account for variability in detection limits and the difference in numbers of PFAS included in each method (Tables S7-8, S10-11). Across all studies, the probability of detecting one PFAS was approximately 18%, with a marked decrease in probability with increasing number of detected compounds; no differences were observed between public-supply and private-well samples (Fig. 3).

Due in large part to funding constraints and prioritization of population-relevant reconnaissance of a range of exposure points within a given community, point-of-use tapwater exposure studies conducted to date (Bradley et al., 2020, Bradley et al., 2018, Bradley et al., 2023a, Bradley et al., 2021a, Bradley et al., 2021b, Bradley et al., 2023b) typically have employed a one-time spatial–synoptic approach, which provides limited to no insight into point-of-use tapwater temporal variability. To address this data gap, PFAS samples were collected temporally at 3 of the residential locations (2 private-wells, 1 public-supply). The rural, private–well location in South Carolina had no PFAS detected during three months of sampling. PFAS were detected at the suburban public-supply and private-well locations in New Jersey, and cumulative detected concentrations were generally stable (hourly increasing to weekly samples) over 3 months (Figures S4, S5). In light of the near-detection-limit concentrations of several individual PFAS detections, the observed variabilities in detections of some individual PFAS and in per-sample cumulative PFAS detections (Figure S4) were likely due more to method–sensitivity limitations (Teymoorian et al., 2023) than to short-term changes in drinking-water resources. New Jersey (NJ), one of the more proactive states regarding PFAS regulation, has established enforceable NJ-MCLs for three PFAS including PFOA (14 ng/L), PFNA (13 ng/L) and PFOS (13 ng/L) (New Jersey Department of Environmental Protection, 2022a) and has added PFOA, PFOS and PFNA to the Private-well Testing Act, which requires homeowners to test potable wells prior to sale and landlords to test their well water and report results to the tenant once every five years (New Jersey Department of Environmental Protection, 2022b). None of the detections in either location exceeded NJ-MCLs. These results further support the need for continued monitoring of PFAS in point-of-use tapwater from residential private-wells as well as for continued temporal assessments to accurately assess PFAS exposures at the point-of-use more broadly.

3.2. Comparison to proposed US drinking water regulations

Newly proposed MCLs for PFOA (4 ng/L) and PFOS (4 ng/L) were released in March 2023 by EPA as part of the National Primary Drinking-water Standards Rule (U.S. Environmental Protection Agency, 2023b). The proposed MCL for PFOS was below the reporting limit for two of the three laboratories (CSM:1.3 ng/L, NERL: 5.0 ng/L and USGS: 7.4 ng/L; Table S3) used during this study; consequently, our estimates of samples exceeding the proposed MCL should be considered conservative. The reporting limit for PFOA was above the proposed MCL for only one laboratory (CSM: 1.3 ng/L, NERL: 5 ng/L, USGS: 2.0 ng/L; Table S3) in which the least number of samples were analyzed (26/716; Table S1). Proposed MCLs for PFOA and PFOS were exceeded in 6.7% and 4.2%, respectively, of all tapwater samples collected but were exceeded in 48% (48 of 99) and 70% (30 of 43), respectively, of tapwater samples when detected. Further, proposed MCLs for PFOA and PFOS were exceeded in 63% and 67%, respectively of the private-well tapwater samples and in 44% and 77%, respectively of the public-supply tapwater samples, when detected. The proposed MCLG (zero) (U.S. Environmental Protection Agency, 2023b) was de facto exceeded in every sample in which PFOS and PFOA was detected (private-well: 15 and 24, respectively; public-supply: 28 and 75, respectively). Further, to account for dose additive noncancer effects of PFBS, PFNA, PFHxS, GenX, EPA proposed an MCL for the mixture of these four PFAS based on a hazard index approach (U.S. Environmental Protection Agency, 2020). The proposed hazard index of 1 for the sum of the toxicity quotient (measured concentration/health-based value) for PFBS + PFNA + PFHxS + GenX was exceeded in 4.6% of tapwater collected.

3.3. Aggregated screening-level assessments

We also used two bioactivity weighted screening approaches Σ_EAR and Σ_TQ to provide insight into the potential aggregated (sum of all PFAS) effects. These approaches are limited by the availability of weighting factors (ToxCast ACC and health-based benchmarks, respectively) and mixture effects are estimated by assuming concentration addition (Cedergreen, 2014). The Σ_EAR approach has been used effectively in other studies as a protective (conservative) screening tool to asses drinking-water exposure risk to organic contaminant mixtures because it leverages response relations for>9000 organic chemicals across over 1000 standardized vertebrate cell lines (Kavlock et al., 2012, Kavlock et al., 2008, Richard et al., 2016). Contaminant bioactivity ratios were aggregated across all ToxCast endpoints available for individual PFAS without restriction to recognized modes of action to provide a precautionary lower-bound estimate of in vivo adverse-effect levels (Paul Friedman et al., 2020), however, this approach may not accurately reflect apical effects (Blackwell et al., 2017, Schroeder et al., 2016). Further, for PFAS, the approach has the potential to underestimate exposure risk because only 11 of the detected compounds had exact Chemical Abstract Services number matches in the ToxCast™ database and only six (PFOS, PFOA, PFNA, PFHpA, PFBA, PFHxS) of these had EAR exceeding our lowest threshold for consideration (>0.00001). The ∑_TQ approach targets apical human-health effects, is notably constrained to recognized (i.e., benchmarked) health concerns and was used herein to estimate cumulative effects from a broader suite of PFAS (15 with established health-based benchmarks; Table S5) (Interstate Technology Regulatory Council, 2022). Lastly, it is important to note, the EAR approach is based on measured endpoint-specific activity cutoff concentrations, whereas the human-health benchmarks used in the TQ approach generally include a margin of safety (margin of exposure).

None of the samples exceeded a Σ_EAR > 1, a value which indicates cumulative exposure at concentrations capable of modulating molecular endpoints in vitro, and we observed no systematic differences in Σ_EAR between private–well and public–supply locations (Fig. 4). However, 65 tapwater samples exceeded the Σ_EAR = 0.001 precautionary screening level of potential concern (yellow line, Fig. 4; Table S10) for molecular effects described previously (Bradley et al., 2018). For PFAS and PFOA, which proposed MCLGs of zero, EPA’s interim health advisory levels released in 2022 (U.S. Environmental Protection Agency, 2022a) were used as the benchmark value (Table S5). Σ_TQ values were higher in private–well tapwater samples compared to public–supply (p = 0.0015), and 124 samples overall had Σ_TQ > 1 (Fig. 4, Table S10), indicating a high probability of aggregated risk when considering exposures to all observed PFAS with an available benchmark. As expected, given their proposed MCLG of zero (U.S. Environmental Protection Agency, 2023b), Σ_TQ results were driven by PFOA and PFOS, when detected. Simultaneous detection of multiple PFAS is consistent with other drinking-water (Andrews and Naidenko, 2020, Boone et al., 2019, Hu et al., 2016) and tapwater studies (Bradley et al., 2020, Bradley et al., 2018, Bradley et al., 2021b) in the US. The results of precautionary Σ_EAR and Σ_TQ assessments by this group indicate that the potential for human-health effects from contaminant exposures (including PFAS) through drinking-water are common and comparable in private-well and public-supply tapwater (Bradley et al., 2020, Bradley et al., 2018, Bradley et al., 2021a). Information generated by this study and elsewhere indicates the need for further assessments of cumulative health exposure risks of PFAS mixtures (Kwiatkowski et al., 2020) and of PFAS in combination with other organic and inorganic contaminants of concern, particularly in unmonitored/unregulated private-wells where information is limited or not available.

3.4. Predicting exposure based on potential sources and land-use

The final model comparing tapwater PFAS concentrations to geospatial drivers included median distance to potential source (including airport, industry, waste management, oil and gas, and department of defense) and several land-use classifications (total developed, open developed, cultivated and pasture agriculture). The number of PFAS sources in a 5–km buffer around each site was highly correlated with developed land-use (r = 0.66) and was excluded from the model. Cumulative PFAS concentrations and the number of detected compounds increased with surrounding developed–land and decreased with increasing distance from probable source(s) (Fig. 5, Figure S6, Tables S10-11). However, for individual PFAS the distance to probable source(s) was not a strong predictor of concentration (Figure S7, Table S8). PFBS was the only PFAS which exhibited a positive relation with development and pasture agriculture and a negative relation with open development (commonly includes large-lot single-family housing units, parks, golf courses, and vegetation planted in developed settings for recreation, erosion control, or aesthetic purposes). Although biosolids application is a recognized source of PFAS to surface- and groundwaters in agricultural landscapes (Munoz et al., 2022, Sepulvado et al., 2011), in this study concentrations of PFBS, PFHxA, PFHxS, PFOA and PFOS decreased with increasing surrounding cultivated cropland. Further, for several compounds (PFBA, PFHxA, PFHxS and PFOA), concentrations decreased with increasing development, an unexpected result which suggests that point-of-use tapwater exposure to individual PFAS in our study is more closely associated with the type of PFAS sources (e.g., industry, airport, wastewater, etc., which are generally located at the edge of urban development and not in the highest residentially-developed areas) (Hu et al., 2016) rather than numbers of potential PFAS sources (Salvatore et al., 2022). Because our geospatial analysis was constrained to broad putative-source categories (e.g., military fire training areas, industrial, wastewater) and 90% of the corresponding potential sources were characterized as industrial, exploration of source-type/point-of-use tapwater exposure relationships was limited and merits further investigation. Further, most public-supply samples were collected at the tap not the treatment facility, indicating a distinct disconnect from PFAS source and drinking water resource. Despite these limitations, the real–world point-of-use tapwater PFAS concentration data compiled herein along a national gradient of presumptive contamination represents an important validation dataset to assess and tune putative-source PFAS contamination models, like that presented recently (Salvatore et al., 2022).

3.5. Implications and future directions

Approximately 40 million people in the US rely on private-wells for drinking-water (DeSimone et al., 2015, Dieter et al., 2018), most national testing programs, like the UCMR3 focused on community water supplies serving ≥ 10,000 consumers, do not include private-wells and rarely capture information from rural communities (52 million people rely on small water supplies serving < 10,000), indicating data on PFAS exposure and potential human-health effects is does not exist for over one-third of the US population (Hu et al., 2016). As noted previously, small public supplies and private-wells may be disproportionally affected by PFAS, emphasizing the value of studies like these focused broadly on point-of-use tapwater PFAS exposures, with an emphasis on comparing exposures in private-wells with those directly from public-supply using similar sampling/analytical methods. Some of these gaps associated with PFAS in small community public supply facilities may be addressed by the UCMR5, currently underway in the US and expected to provide extensive information on PFAS in US drinking water for public-supply consumers in the next few years (U.S. Environmental Protection Agency, 2023a).

Modeled results indicate that on average at least one PFAS is detected in about 45% of US drinking-water samples. Results also indicate that 1) detection probabilities vary spatially (8% in rural areas up to > 70% in urban areas/areas with a known history of PFAS contamination), 2) drinking-water exposures may be more common in the Great Plains, Great Lakes, Eastern Seaboard, and Central/Southern California regions, and 3) temporal variations in concentrations/detections may be limited. Geospatial datasets and land-use information were correlated with both cumulative PFAS concentrations, and the number of PFAS detected; however, they were not often correlated with specific PFAS profiles due to the limited number of individual PFAS detected more than once using targeted approaches. Targeted PFAS analytes are only a fractional indicator of the 8000 + potential PFAS and the fraction of total organic fluorine captured by these targeted analyses is typically low in surface water (D’Agostino and Mabury, 2017, McDonough et al., 2019) and drinking-water (Jiao et al., 2022). Potential detection of one or more PFAS in US drinking-water combined with the paucity of information available on current use/ultra-short chain compounds (Neuwald et al., 2022) supports the continued need for point-of-use tapwater monitoring, with an emphasis on unmonitored private-wells and underserved communities on small community water supplies. To fully understand exposure and adequately determine risk to human-health, continued emphasis should be placed on 1) integrating geospatial datasets with PFAS data broadly to identify vulnerable regions/subpopulations, 2) expanding monitoring to include rural small–system and private–well dependent communities, and 3) expanding target and non-target analysis methods particularly in drinking-water monitoring programs in the US and globally.