Inorganic arsenic is a known human carcinogen causally associated with cancers of the skin, bladder, and lungs, and epidemiologic evidence supports a potential association with cancers of the kidney, breast, pancreas, and liver [11, 12]. Chronic exposure is also associated with respiratory disease, cardiovascular disease, adverse birth outcomes, metabolic disorders and diabetes, impaired immunological functioning, kidney disease, and adverse neurocognitive outcomes [13, 14]. Risk of some outcomes including cancer likely persist for chronic exposure to water arsenic at concentrations at or below EPA’s MCL of 10 μg/L [15, 16].

The significant spatial variability in water arsenic concentrations across the US reflects variability in geologic, biogeochemical, hydrologic, and climatic conditions that influence geogenic arsenic prevalence in bedrock and solubility, (e.g., arid oxidizing environments in the Southwestern US, and humid reducing environments and alkaline pH in the Northeast) [17]. Arsenic is detectable in over 50% of CWSs contributing to the EPA’s Six Year Review database (>36,000 CWSs) [18]. Approximately 2.6% of CWSs report arsenic concentrations exceeding the MCL (10 μg/L). An estimated 85% of PWSs rely on groundwater sources. Predicted arsenic levels in domestic wells are strongly associated with CWS concentrations using the same groundwater resources [19]. Hardrock mining processes and mine waste also contribute to surface and groundwater arsenic in the Southwest and Great Plains regions of the US, especially near Indigenous communities [20].

Significant sociodemographic and regional inequalities in CWS arsenic exposures have been identified across the US. In 2009–2011, mean arsenic concentrations were 1.70 μg/L nationwide and were highest in: systems serving communities categorized as Semi-Urban, Hispanic (3.40 μg/L); communities in the Southwestern US (3.18 μg/L); communities with less than 500 residents; communities reliant on groundwater sources; and incarcerated populations in the Southwest (2006–2011, mean 6.41 μg/L) [18]. At the county level, a higher proportion of Indigenous or Hispanic/Latino residents was associated with higher CWS arsenic concentrations, after adjustment for socioeconomic vulnerability [21]. A California study similarly found that higher proportions of Latino residents were associated with significantly higher CWS arsenic concentrations [22]. Similarly, a higher proportion of non-Hispanic Black residents was also associated with higher CWS arsenic concentrations in the Southwestern US, while concentrations were inversely associated with higher proportions of non-Hispanic White residents both nationwide and regionally. Higher county-level high school diploma attainment was associated with lower CWS arsenic concentrations; this association was modified by region, likely reflecting regional/local differences in socioeconomic context and drinking water infrastructure [23].

Evidence from EPA violation records, routine compliance monitoring records, and urinary biomarkers of water arsenic exposure indicates that public water arsenic exposure declined significantly following the reduction of EPA’s MCL from 50 to 10 μg/L (effective 2006), consistent with the WHO guideline of 10 μg/L [18, 24–26]. The largest exposure reductions occurred for Mexican-American residents (36% reduction, compared to 17% overall) and for CWSs with the highest baseline arsenic concentrations, including those serving New England (mean 37% reduction) and Alaska and Hawaii (mean 24% reduction).

EPA’s MCLG of 0 μg/L for arsenic reflects that there is no safe level of exposure to carcinogens, and mounting epidemiologic evidence supports that the current MCL is inadequate to protect human health. Denmark and the states of New Jersey and New Hampshire set more health protective MCLs of 5 μg/L, and water utilities in the Netherlands voluntarily adopted 1 μg/L. In setting the 5 μg/L MCL, the New Jersey Department of Environmental Protection (NJDEP) cited the National Academy of Sciences’ 2001 report that a water arsenic concentration of 0.003 μg/L is estimated to result in a one-in-one-million excess lifetime risk of lung and bladder cancer [27]. NJDEP considered available testing and treatment technologies but not cost-benefit analysis [22]. Significant uncertainties and a lack of overall scientific consensus remain regarding the risk assessment for inorganic arsenic. The current EPA Integrated Risk Information System (IRIS) cancer slope factor for arsenic (1.5 per mg/kg bodyweight-day) relates to skin cancer only, while the 2010 proposed slope factor (25.7 per mg/kg bodyweight-day) corresponds to combined lung and bladder cancers and reflects increased susceptibility for women; synergistic effects for tobacco smoking are not considered in either [28].

Climate change poses significant challenges to reducing water arsenic concentrations, especially in the Southwest where water arsenic levels are already high. Wildfires and extended drought conditions caused by climate change are likely to concentrate arsenic and other inorganic contaminants as water levels decrease in groundwater [19, 29]. Additional epidemiologic studies of drinking water arsenic at low- to moderate levels relevant for US populations, especially investigating cancer and cardiovascular disease, in diverse US populations would further inform risk assessment efforts.

Disinfection of drinking water is necessary to prevent waterborne infections. However, disinfectants are highly reactive and interact with organic matter, bromide, nitrogenous compounds and other precursors to form unintended disinfection by-products (DBPs) [30, 31]. DBP concentrations can vary substantially based on source water, disinfectant and treatment practice. Chlorine is a cost-effective disinfectant widely used worldwide, and trihalomethanes (THMs) followed by haloacetic acids (HAAs) are the most prevalent chlorination by-products. THMs comprise chloroform, dibromo-chloromethane, bromodichloromethane, and bromoform [32]. Chloroform is reported to be the dominant THM (up to 90%) in many areas worldwide, while brominated THMs are the more abundant species elsewhere, such as in Middle East countries, associated with high concentrations of bromide ions in raw water [33]. Local bromide discharges from industrial sources, including coal-fired power plants, oil and gas extraction activities, and textile mills, will impact DBP risks [34].

Alternative disinfectants may decrease THM formation but may encourage formation of other DBPs, such as chlorate and chlorite (from chlorine dioxide), nitrogenous DBPs (chloramines), bromate (ozone), iodinated and brominated aromatic DBP (chlorine dioxide). Chloramines are widely used in the US as an alternative to chlorine to reduce THM formation. Some of these DBPs show higher toxicity at lower concentrations than THMs and HAAs [33]. Among the ~700 identified DBPs, only 4 THMs, 5 HAAs, bromate, chlorate, and chlorite are currently regulated in the US and/or the European Union (EU) (Table 3) [35, 36].

Long-term exposure to THMs, as a marker of DBP exposure, has been consistently associated with bladder cancer risk in epidemiological studies [37]. DBP exposure has also been linked to pregnancy and reproductive outcomes, but evidence is mixed [38, 39]. The WHO International Agency for Research on Cancer has classified chloroform and bromodichloromethane as possible human carcinogens [40]. Iodinated DBPs are more toxic in mammalian cells than their chlorinated and brominated analogues [41]. Iodinated DBPs may be among the most genotoxic and cytotoxic DBPs, with iodoacetic acids potentially the most genotoxic of all DBPs [42]. Endocrine disruption and adverse reproductive and developmental impacts are also seen [43–45]. Some nitrogenous DBPs, including haloacetonitriles, haloaceta-mides, and halonitromethanes are more toxic and carcinogenic than THMs and HAA [43, 46]. In particular, the toxicity of haloacetamides is estimated to be 142 times higher than HAAs [47].

DBP formation generally is higher in surface than groundwater systems due to higher levels of natural organic matter. The composition of organic matter and other water constituents in raw water affects which DBPs are formed. In particular, hydrophobic organic matter (e.g., high molecular weight organic materials) has higher potential to form THMs, nitrogenous DBPs and aromatic DBPs than hydrophilic organic matter. Occurrence of bromide and iodide ions promotes the formation of brominated and iodinated DBPs, respectively, while ammonia in water favors nitrogenous DBP formation [46].

Seasonal fluctuations (involving variations of surface water quality and water temperature) influence DBP formation. DBPs tend to show higher concentrations in summer than winter and certain DBP groups exhibit distinct seasonal patterns. Hydrophobic natural organic matter is positively correlated with air temperature whereas the hydrophilic natural organic matter shows a reverse trend [48]. Temperature increases the reaction rate between organic matter and chlorine, increasing THM formation. HAA formation, however, increases with temperature up to 20 °C then decreases. High temperatures also increase THM volatilization. Likewise, high pH enhances THM formation as well as the hydrolysis of other DBPs into THMs and HAAs, while low pH increases HAA formation [33, 49, 50].

THM and HAA formation increases with disinfectant dose and residence time in the distribution system, so distal parts of the distribution system generally have higher DBP levels. DBP exceedances occur more often in small systems, which have fewer resources for treatment to reduce DBP levels, such as filtering organic matter prior to disinfection [51]. DBP formation in the distribution system may be higher with polyethylene pipes [52] and with increasing diameter and pipe age [53].

Tap water is also used for showering, bathing, dishwashing and cooking. While exposure to non-volatile DBPs occurs predominantly through ingestion, exposure to volatile and skin-permeable DBPs (e.g., THMs) also occurs through dermal absorption and inhalation [32, 54, 55]. Exposure to THMs in swimming pools has been evaluated (measuring THMs in blood, exhaled air and urine), but the relative importance of different exposure routes remains inconclusive [56].

The pervasive presence of DBPs in drinking water poses significant concerns for human health. Even regulated DBPs lack comprehensive toxicological evidence, and while data indicate that some unregulated DBPs may pose greater risks that those currently regulated, the majority of emerging DBPs remain poorly understood. Furthermore, the potential interactions among the 700 identified DBPs in drinking water, have not been adequately examined individually or in mixtures. Synergistic effects of climate change (e.g., increasing temperature, more frequent and severe flooding and droughts), acidification of soil and surface water, land use change and other anthropogenic pressures all impact water quality and consequently, water treatment and DBP formation [57]. Precipitation and temperature are the main climate factors affecting water quality. The increase in atmospheric temperature and warming of surface waters is linked to eutrophication and increased microbial activity and dissolved organic carbon. Drought has been shown to significantly impact water chemistry, including higher levels of hydrophilic organic matter [58]. Ultimately, all these factors impact the concentration and composition of organic matter and consequently, DBP formation [57]. Aging and outdated water infrastructure and the large number of small systems are also challenges.

Unconventional oil and gas development (UOGD), commonly called “fracking”, is a method for extracting oil and natural gas from deep, low permeable geologic formations, requiring more intense stimulation compared to more accessible, conventional reservoirs [59]. In the US, there are approximately 150,000 active UOG wells, and more than 9 million people rely on drinking-water sources located within 1.6 km (1 mile) of a UOG well [59]. Water contamination from UOGD remains a major community concern [60].

Fracturing fluids and wastewater used or produced by UOGD may contain toxic, radioactive, endocrine-disrupting, and/or carcinogenic chemicals [6, 61]. Potential water contamination events include surface spills of fracturing fluids or wastewater at the well site, release of improperly treated wastewater, and leaks in well infrastructure [62]. An estimated 1–4% of UOG wells have reported spills [63, 64] and, based on Pennsylvania data, approximately 20% have a non-administrative violation [65]. These are uncommon events, and multiple groundwater monitoring studies conducted in regions with UOGD have not found evidence of widespread contamination [66–69]. However, specific instances of groundwater and surface water impairments have been identified [70–72].

Numerous epidemiologic studies have observed an increased risk of adverse health effects including adverse perinatal outcomes, childhood cancer incidence, hospitalizations, asthma exacerbations, mental health issues, and mortality in the elderly in relation to proximity to UOGD sites [59]. The extent to which these associations may be related to water is unclear, because most studies have used aggregate proximity-based metrics to assign exposures, which are not specific to any hazard [59]. A few epidemiologic studies have focused specifically on the drinking water exposure pathway. Two studies of pediatric health outcomes applied a novel water-pathway specific metric that restricts the analysis to UOGD wells that are hydrologically connected to the watershed of a residence [66, 73]. This exposure metric is most relevant for groundwater. Another study found that proximity of community drinking water sources to UOGD wells was associated with greater likelihood of a variety of adverse birth outcomes [74, 75]. More research is needed to understand whether the increased health risks observed in populations living in the vicinity of UOGD are attributable to drinking-water exposures, other hazards, or a combination of factors.

In the US, UOGD often occurs in rural areas where homes are served by private (domestic) drinking water wells, which are not subject to federal regulations and monitoring [76], or by small water systems. Therefore, available data are quite limited, and the data that are available tend to focus on the few chemicals that are regulated, which are only a tiny subset of the approximately 1200 chemicals reportedly used or produced by UOGD.

Although chemical disclosure policies have improved, researchers still lack a complete and consistent inventory of chemicals used in UOGD [77]. Furthermore, contaminants also include transformation products and naturally occurring chemicals mobilized during the development process [78]. Some researchers have mined existing voluntary reporting databases and created their own datasets [79].

Proximity to UOGD wells, often emphasized for sampling and in risk estimation, may be an inadequate surrogate for predicting exposure given geologic heterogeneity and topographical variations producing the complex flow paths and stochastic nature of contamination events [66]. Applying physically based hydrological models demonstrates that unlike the typical circular buffers used in exposure and epidemiologic studies, the groundwater capture zones exhibit more of a surfboard shape [80]; better modeling the capture zone could improve identification of homes more vulnerable to contamination and inform sampling locations. In addition, application of machine learning techniques to available monitoring data can help identify hotspots [81, 82].

Finally, information on the locations and descriptions of where violations and spills occur is not available in real time and therefore timely sampling in response to a potential contamination event is not feasible. In addition, violations data are not available in a consistent format across states, posing another challenge to multi-state research. Further, chemicals disperse at varying rates (particularly in groundwater), so a single collected sample may be unlikely to coincide with release of a plume of multiple contaminants, potentially necessitating repeated measures.

Lead is a widely used element with thousands of applications. So closely associated with the conveyance of water, the very word ‘plumbing’ derives from its Latin name, plumbum. Lead is also highly toxic and associated with adverse health endpoints across virtually all body systems, including nervous, cardiovascular, renal, immunological, hematological and reproductive/developmental systems in men and women, in adults and children [83–85]. Lead has been classified as a probable human carcinogen by EPA since 1988 based on rodent toxicology data [83].

Lead is usually a corrosion by-product in drinking water related to water’s natural corrosivity and lead’s extensive use in plumbing components such as pipes, solder, brass and bronze [86], faucets [87, 88], galvanized steel pipe coatings [89, 90], valves and meters. Lead has been progressively restricted from plumbing use in the US during the past few decades, but its durability means that an estimated 9.2 million lead pipes installed in the late 1800’s to early 1900’s remain servicing US homes [91, 92].And until banned in 1986 [93], lead solder joining copper pipes was practically ubiquitous in the US.

Studies evaluating multiple lead exposure sources from within the house found that, when present, lead pipes contribute the most lead to drinking water [94]. Lead pipes carrying water from the water main to the residence (lead service lines) and lead pipes inside homes were often installed in cities that expanded greatly during the Industrial Revolution [95] therefore home age and the history of the urban area are factors to consider. However, lead pipes continued to be installed in the US until banned in 1986, including locales such as Chicago that required lead service line installation until then. Beginning in the 1930s, copper pipes replaced lead pipes as the most common residential piping material; lead solder was used to join them. The combination of copper and lead produces galvanic corrosion that is associated with high lead leaching potential and elevated water lead levels (WLLs).

All water is corrosive, but the degree of corrosivity varies. Principal factors include pH, alkalinity and hardness of the water [96]. Seasonality is evident in WLLs [97, 98]. The warmer the temperature the greater the potential for lead leaching; consumption of drinking water also increases in the summer.

EPA issued the Lead and Copper Rule (LCR) in 1991. Lead contamination of tap water relates mostly to water corrosivity and corrosion within the lead service line and residence; and water lead levels vary due to stagnation time, temperature, extent of lead plumbing components, and other factors. Citing some of these factors, EPA decided not to set an MCL for lead. EPA’s Lead and Copper Rule (LCR), instead, established a Treatment Technique requiring systems to control the corrosiveness of their water. If more than 10% of the tap water samples exceed the Action Level (AL) of 0.015 mg/L, water systems must take additional steps. The AL was based upon feasibility; it is not enforceable and exceeding it is not a violation of the SDWA. In January 2021, EPA revised the LCR [99] and almost immediately agreed to review it to address shortcomings [100]. The MCLG is zero based upon both lead’s carcinogenicity and that no safe level of exposure has been determined.

The SDWA only regulates PWSs. Homes with private wells, which are not covered under the SDWA, are less likely to have lead pipes than older urban areas. On the other hand, they may be more vulnerable to use of leaded solder and are less likely to use a corrosion inhibitor even with very acidic water; they may also have lead in their water pumps [101]. One study found that WLLs from private wells may be higher than those in adjacent public water systems [102].

Numerous studies show the clinical significance of exposures to even low WLLs across a range of effects. WLLs are associated with cognitive performance in children [103], renal function in dialysis patients [104], adverse birth outcomes [105], iron deficiency in patients with End Stage Kidney Disease [106], and the likelihood of juvenile delinquency [107]. Water-lead outbreaks are attested to raise general population BLLs [108, 109] and BLLs are also associated with WLLs in non-crisis circumstances [110]. Violations of the LCR are common. The Government Accountability Office (GAO) reported that at least 10% of the water systems subject to the LCR had at least one open violation of the rule [111]. This is likely an underestimate as GAO has repeatedly found that EPA’s enforcement data are incomplete, especially related to compliance with the LCR [111–113]. This is consistent with EPA’s finding that only 8% of LCR treatment technique violations reported to states are passed on to EPA. Similarly, a 2016 report found that over 5300 CWSs serving an estimated 18 million Americans violated the LCR that year [114]. A later study found that 186 million Americans receive water from CWSs with detectable lead contamination [115]. A familiar pattern is evident in violations of the LCR in the US: minoritized and low-income communities bear an increased risk of receiving poorer quality drinking water and of having lead pipes [116–118].

Climate change effects include the acidification of the natural world, reducing the pH in air, water and soil. Lower pH is associated with increased lead mobility and bioavailability [119–121]. Hence, climate change and global warming may increase water lead levels both through thermal and biochemical mechanisms as well as increasing lead’s mobility and bioavailability. Global warming will likely increase water consumption, also.

Disparities exist in lead exposure, enforcement issues and problems with school drinking water. US lead exposures from drinking water appear to be widely underestimated related to systematic poor monitoring, reporting and enforcement [122]. Less data are available on lead contamination of drinking water outside the US, but there is evidence of underestimation of lead contamination in EU drinking water, also [123]. The WHO provisional guideline value is 10 μg/l, set in 2011; sampling protocols differ between the US and WHO [124]. Until the remaining 9.2 million lead pipes are replaced and effective corrosion control is widely adopted, drinking water will remain a significant lead exposure source.

Nitrate levels in water resources have increased world-wide from applications of inorganic fertilizer and animal manure in agricultural areas [125]. Contamination sources also include septic systems that do not effectively remove nitrogen and discharges from wastewater treatment plants [126], as well as atmospheric deposition of nitrogen oxides and fertilizer use on lawns, golf courses, and parks.

Private wells typically have higher nitrate concentrations than CWSs due to their shallower depth. High nitrate concentrations (near/exceeding the MCL) are most common in shallow (<100 feet) private wells located in agricultural areas because of nearby nitrogen sources (fertilizer use, animal operations, septic systems) [127]. Treatment of private wells is the responsibility of the property owner, leading to racial/ethnic, rural/urban, and socioeconomic disparities in access to safe drinking water. Some states provide resources, such as subsidized water test kits, to private well users. However, state-level regulation of private wells varies dramatically, and private well users are often unaware of the resources available to them [128].

A 2019 study [129] evaluating nitrate exposures in US CWSs estimated that about 5.6 million people were exposed to water with ≥5 mg nitrate-nitrogen (NO3-N)/L (more than half the MCL of 10 mg NO3-N/L) between 2010 and 2014. Hispanic/Latino residents were more likely to be served by CWSs with elevated nitrate levels. Disproportionately high exposure to nitrate-contaminated water among Hispanic/Latino communities has also been identified in the Yakima Valley of Washington State and San Joaquin Valley of California, among other areas [130]. Additional research has documented cases of poor water quality, including high nitrate concentrations, in drinking water used by migrant worker communities, Alaska Native villages and other Tribal lands, and in colonias along the US-Mexico border [130]. Limited measurement data characterizing residents using private wells presents challenges but recent advances in exposure modeling have proved useful for identifying exposure disparities [131–133].

The US recommended standard for nitrate in drinking water was originally set in 1962 by the US Public Health Service as 10 mg NO3-N/L, based on infant methemoglobinemia. EPA’s subsequent MCL only considered this outcome and was based on a no-adverse-effect concentration for drinking water used to prepare formula for infants <6 months of age with no margin of safety; other health effects were not considered [134, 135]. The MCL has not been revised since it was promulgated in 1975 [135]. The literature investigating the health effects of nitrate exposure has expanded greatly since the MCL was set [136]. Nitrate ingested from drinking water may increase the risk of birth defects and some cancers because nitrate is a precursor in the formation of N-nitroso compounds (NOC), many of which are teratogens and carcinogens. NOC are formed in the body (a process called endogenous nitrosation) when nitrate is consumed in the absence of antioxidants that inhibit their formation [137]. Among epidemiologic studies with individual-level data, seven studies (in Australia, Canada, California, Texas, and Denmark) found increased central nervous system (CNS) malformations in children whose mothers consumed drinking water with high nitrate concentrations during pregnancy. In six studies, the increase in CNS malformations occurred at levels below the MCL [136, 138–140]. Studies of spontaneous abortion, fetal growth, and birth weight have been more limited and had mixed results [136, 139, 140].

In 2006, the International Agency for Research on Cancer (IARC) concluded that when ingested under conditions that result in endogenous nitrosation, nitrate and nitrite are probably carcinogenic to humans [137]. Since the IARC review, there have been more than 20 studies of incident cancer mostly in the US and Europe. Colorectal cancer is the most well-studied, with four of five studies finding increased risks [136, 141]. Studies of other incident cancers were fewer; however, positive associations at levels below the MCL were observed for cancers of the bladder [136, 142], kidney [136], childhood and adolescent/young adult brain [143–145], ovary and thyroid [136].

Unlike nitrate in drinking water, nitrate naturally present in food (mostly in fruits and vegetables) is consumed together with antioxidants, vitamins and polyphenols that inhibit endogenous nitrosation [146]. In controlled longitudinal feeding studies, high nitrate intake through consumption of high-nitrate vegetables such as beets and dietary supplements has been shown to reduce hypertension and may play a role in the protective effect of vegetables on cardiovascular disease risk [147, 148]. However, hypertension and nitrate ingestion through drinking water sources has not been studied. Clinical and subclinical hypothyroidism have been linked to higher nitrate intake from drinking water in several studies [136] and deserve further study.

Additional studies of cancers, thyroid disease, and birth outcomes/defects that have shown the most consistent associations with drinking water nitrate are needed to further elucidate risks below the MCL. Evaluating subgroups with higher endogenous nitrosation will improve inference. Methods for quantifying the nitrate-reducing bacteria in the oral microbiome and characterizing genetic variation in N-nitroso compound metabolism hold promise for identifying these high-risk groups in epidemiologic studies.

Over the past several decades, nitrate levels in many ground and surface waters increased despite efforts to reduce nitrogen inputs [127]. Nitrate concentrations are expected to increase in aquifers used for drinking water as the contamination in shallow groundwater moves to deeper aquifers [127]. Disparities in exposure, an outdated health assessment, and large numbers of contaminated small systems and private wells are among the shared challenges related to nitrate. Efforts to understand the disproportionate impacts of nitrate exposure will help inform future policy and regulations to limit sources of nitrate contamination, decrease exposure, and alleviate public health harms.

Per- and polyfluoroalkyl substances (PFAS) are a major class of contaminants of concern in drinking water across the US and globally. PFAS as a class are generally considered “emerging” although some individual chemicals have state-level MCLs or have been phased out of commercial production may be classified as “legacy” pollutants. PFAS exposures are widespread; according to the National Health and Nutrition Examination Survey (NHANES), over 98% of the US population has detectable levels of PFAS in their blood [149]. In areas with contaminated water supplies, drinking water is a major contributor to PFAS exposure, although exposures can also come from diet, consumer products, and building materials. Dubbed “forever chemicals” because of their extreme persistence in the environment, PFAS also raise concerns due to far-reaching toxicity and bioaccumulation potential, with certain long-chain PFAS having human half-lives in blood on the order of years [150]. Exposures to some PFAS have been associated with adverse health outcomes across many major systems in the body, including immunotoxicity, dyslipidemia, changes in thyroid hormone levels, decreased birth weight, and testicular and kidney cancer [151–153]. Much of the available toxicological and epidemiological evidence is based on a relatively small number of long-chain PFAS (often those considered “legacy” PFAS), especially perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA), which have been phased out of production in the US and Europe, but reportedly are still being produced internationally and imported into the US in consumer goods [154]. However, a growing body of evidence raises concerns about the toxicity of “emerging” alternative PFAS being used as replacements [155, 156].

The highest levels of PFAS in drinking water have been found close to industrial facilities where PFAS are manufactured or processed and sites with discharges of aqueous film forming foam (AFFF) at military bases, major airports, and other fire training areas [157]. PFAS are widely used in consumer items, such as stain-resistant carpets and upholstery, food packaging, apparel, and cosmetics. An increasing number of PFAS contamination sites have been linked to waste disposal, including land-applied biosolids, effluent from wastewater treatment plants and septic systems, and landfill leachate [158].

PFAS have increasingly been detected in PWSs as analytical sensitivity has improved and testing has become more wide-spread. In the third round of EPA’s Unregulated Contaminant Monitoring Rule (UCMR3) in 2013–2015, which included six (mainly legacy) PFAS compounds, only 4% of PWSs reported detections above minimum reporting limits (MRLs) [157, 159]. However, this testing greatly underestimated the extent of PFAS in PWSs as the MRLs were relatively high (10–90 ng/L) and most smaller PWSs (≤10,000 customers) were not included. A more recent analysis estimated that 18–80 million Americans are served by PWSs delivering ≥10 ng/L of PFOS and PFOA and that the water of 200 million Americans has concentrations of PFOS and PFOA ≥ 1 ng/L [160, 161]. In 2023, EPA estimated that from 70 to 94 million people in the US are exposed to six PFAS of concern in their drinking water at elevated levels [162]. Private wells also can be vulnerable to PFAS contamination, even in areas without industrial activity [126] but little testing has been conducted on private wells [163]. PFAS contamination of drinking water supplies is emerging as an environmental justice concern. A recent analysis of monitoring data from 7873 CWSs in 18 US states found that detection of several PFAS is positively associated with the number of PFAS sources and proportions of people of color (Hispanic/Latino, non-Hispanic Black) who are served by these water systems. There are also disparities in the extent of PFAS testing; a smaller proportion of Tribal PWSs were included in the UCMR 3 testing compared to non-Tribal PWSs, and this difference will likely persist in the UCMR 5 testing currently underway [164].

Understanding the characteristics of PWSs most likely to have PFAS can help prioritize PFAS testing in areas where drinking water is most vulnerable. In the UCMR 3 testing, detection frequencies were twice as high among PWSs that relied on groundwater sources compared to surface water, although this testing found that short-chain PFAS were more frequently found in surface water systems [157]. These patterns vary in different regions of the world; nationwide testing of Swedish drinking water found that detection frequencies in surface water systems were twice as high as for groundwater systems [165]. In a 2022 study of groundwater in the eastern US, PFAS were detected in 60% of public supply wells and 20% of private wells, and PFAS detections were correlated with nearby urban land use, tritium (a marker of recent recharge), volatile organic compounds, and pharmaceuticals [166].

In the absence of enforceable federal drinking water standards, a regulatory patchwork emerged as some individual states adopted their own regulations. In 2016, EPA lowered its non-enforceable health advisories for PFOA and PFOS from 400 ng/L and 200 ng/L, respectively, to 70 ng/L for the two compounds individually or combined. Between 2016 and 2022, 18 US states adopted health advisories or enforceable standards at levels lower than 70 ng/L for PFOA and PFOS and/or for other PFAS compounds, mainly in the range of 10–20 ng/L for several PFAS compounds, individually or combined [167]. In 2022, EPA issued drastically stricter interim health advisories of 0.004 ng/L for PFOA and 0.02 ng/L for PFOS, noting “the levels at which negative health outcomes could occur are much lower than previously understood,” and finalized health advisories for two other PFAS [168]. In March 2023, EPA issued long-awaited draft MCLs of 4 ng/L for both PFOA and PFOS (individually), reflecting the “lowest feasible quantitation level.” [162] At the same time, EPA also issued a third MCL for four additional PFAS (PFHxS, PFNA, PFBS, and GenX chemicals) using a Hazard Index value of 1. The Hazard Index approach is routinely applied in risk assessment settings, but this represents the first proposed use for setting drinking water MCLs and represents a step towards more of a class-based approach by moving beyond a one-at-a-time approach to limiting PFAS in drinking water.

PFAS pose numerous challenges for drinking water providers and regulators. With an estimated 14,700 compounds classified as PFAS [7], the full extent of PFAS in water is likely underestimated by current analytical methods, which typically target only 20–30 compounds. The fifth cycle of UCMR (UCMR 5) will include 29 PFAS, including a range of both legacy and newer alternative PFAS [169]. Methods to estimate total PFAS (e.g., extractable organofluorine) or that target certain precursor compounds (e.g., total oxidizable precursor assay) are not widely applied to drinking water. Although thousands of water systems have discovered PFAS contamination, the full extent is unknown as testing has been inconsistent [160].

The cost of PFAS monitoring and treatment itself places substantial financial burdens on PWSs, especially those serving small and low-income communities. For example, the Hyannis Water System in Barnstable, MA, which serves 14,000 customers and includes environmental justice neighborhoods, has spent over $20 million to install granular activated carbon (GAC) treatment on 11 groundwater wells and will incur annual operating costs of $800,000 [170]. The cost for some large water systems could exceed $1 billion over time [171]. The most common PFAS treatment methods used by PWSs (GAC, ion exchange, reverse osmosis) are non-destructive, creating substantial quantities of contaminated filter media and wastes. High-temperature incineration under carefully-controlled monitoring and conditions has been reported in limited studies to largely break down PFAS and regenerate GAC filter media for reuse [172]. However, additional in-depth study is needed to confirm the efficacy of incineration/regeneration at fully destroying PFAS in spent GAC media under a variety of conditions, and use of such incineration has raised concerns about the sustainability of long-distance transport of spent GAC and air emissions from incineration potentially exacerbating exposures in environmental justice communities. New destruction methods such as super critical water oxidation are currently not commercially available, but in the future may provide a path forward to treat contaminated media and waste [173].

Many scientists, regulators, and advocates support class-based approaches to restricting PFAS [174]. Others, including industry representatives, have argued that applying a class-based approach to PFAS in drinking water is complicated by a lack of toxicity data for many PFAS and the different potencies of individual PFAS compounds. Shared challenges also include the lack of health assessments and occurrence data for the enormous class of PFAS that continues to grow literally daily. Finally, addressing PFAS contamination will require a concerted effort to limit PFAS manufacturing to avoid new sources of PFAS and new regulatory approaches that assess individual and combinations of PFAS, enhanced and more widespread testing to understand the true scope of water contamination, and development of new remediation methods to degrade PFAS without creating new exposure risks.

Uranium (U) occurs naturally in the earth’s crust, with water contamination resulting from geochemical processes. Exposure to uranium in the US population is widespread; 74% of NHANES participants from 2001 to 2010 had detectable concentrations of U in their urine. Urine is the preferred biomarker to assess chronic exposure in populations with constant exposures, as previous studies have identified a good correlation between urine U and environmental U in water, air and food [175–178]. Drinking water remains the main route of U exposure in the US [179].

In the human body, uranium is rapidly distributed and accumulates in the bone and the kidneys, which are the main target organs that have been used in determining chemical toxicity for water standards [180]. Alpha radiation from uranium decay is classified as carcinogenic, and increasing epidemiological evidence shows that exposure to uranium in its metallic form is associated with chronic kidney disease, as well as neurologic, reproductive and cardiovascular toxicity [181–183]. While most epidemiologic studies have been conducted in occupational populations with high levels of exposure [184, 185], recent work on community level exposures in Indigenous communities exposed to uranium mine waste has identified more sensitive endpoints including cardiovascular disease and immune dysfunction [186, 187], and is supported by laboratory and animal uranium exposure model studies [188, 189]. Prior development of drinking water standards incorporated the long-held assumption that ingested uranium is very poorly absorbed from the gut. However, more recent studies suggest the higher resulting exposure of regulatory immune cells lining the gut may still result in central dysregulation of the immune system [190]. Thus, additional epidemiological studies are needed to better understand and characterize the adverse health effects of uranium drinking water exposure at the moderate and low exposure levels common in the general population, as well as to identify vulnerable subpopulations.

Uranium mobilization is influenced by the redox environment and increases in oxic groundwaters and in the presence of carbonate complexes, which can lead to the persistence of uranium in drinking water even after treatment to remove chemical contaminants [191, 192]. U is consistently found co-occurring with other metals in groundwater, mostly arsenic and selenium [193–196]. Anthropogenic activities, including mining of uranium ore, producing phosphate fertilizers, and military operations, can lead to increased uranium contamination in drinking water [177, 181, 197]. Depleted uranium (DU), a radioactive byproduct obtained from uranium enrichment and primarily derived from human activities, is one of the contributors to increased U exposure through drinking contaminated water and inhalation of DU aerosols. DU is used for the production of military and hospital equipment and has a half-life of millions of years [177]. Exposure to DU has been associated with renal, neurological and adverse developmental effects in previous observational studies [198, 199]. As a result, U concentrations in water supplies and air are highest in regions with natural geogenic uranium presence and redox conditions that facilitate its release, as well as intense mining, industrial, and military activities.

U is widely detected in private domestic wells and CWSs across the US. According to data from the US National Water Information System (NWIS), 50% of domestic wells in the US have detectable concentrations of U, with ~4% of wells exceeding the EPA MCL of 30 μg/L [200]. Uranium is also detected in 63.1% of regulated CWSs, with ~2% of CWSs exceeding the MCL [193]. Nationwide, CWSs reliant on groundwater have higher mean U concentrations (4.67 μg/L), compared to those reliant on surface water (1.79 μg/L) [193]. Of the 161,000 abandoned hard rock mines in the Western US states, U was the second highest prevalence of primary ore mined, creating the expectation of regionally higher concentrations of uranium in those regions, regardless of surface or groundwater sources, resulting from both anthropogenic contamination and natural mineralization [20].

Several sociodemographic and geographical inequalities in U concentrations at the CWS level have been documented. Nationwide, the mean concentration of uranium in 2000–2011 was 4.4 μg/L [193]. Higher mean concentrations of uranium were detected in CWSs serving populations less than 500 (5.04 μg/L), those serving communities characterized as “Semi-Urban Hispanic” (10.04 μg/L), and CWSs serving communities in the Central Midwest (8.04 μg/L) and Southwest (9.13 μg/L) regions [193]. At the county level, a recent nationwide study identified that higher proportions of residents who self-identify as Hispanic/Latino, American Indian and Alaskan Native were associated with higher uranium in CWSs, after adjustment for income and education [201]. Previous studies in the Navajo Nation have documented higher concentrations of uranium and other toxic metals in drinking water sources compared to the rest of the US [194–196]. At least 12% of unregulated water sources in the Navajo Nation have uranium levels above the MCL of 30 μg/L [202]. Consistent with the exposure data, epidemiological studies have identified 2–3-fold higher urinary levels of uranium among Navajo people compared to the general population levels documented in NHANES [203], and urinary uranium concentration among pregnant women living on Navajo Nation in the Southwest are 2.67 to 2.80 times higher compared to the general US population [156]. Exposures related to private wells are a key challenge.