IntroductionPer- and polyfluoroalkyl substances (PFAS) present a unique challenge in water and wastewater treatment (Vo et al. 2020). Conventional physicochemical and biological treatment methods cannot fully degrade or mineralize PFASs in drinking water treatment (Belkouteb et al. 2020; Kim et al. 2020; Boone et al. 2019; Dauchy 2019; Page et al. 2019; Vughs et al. 2019; Hopkins et al. 2018) or wastewater treatment (Chen et al. 2018; Szabo et al. 2018; Arvaniti and Stasinakis 2015) and may generate PFAS transformation products. The treatment goal of mineralization relative to PFAS is defined by Horst et al. (2020) as complete defluorination regardless of whether the carbon is fully oxidized to carbon dioxide. In contrast to conventional treatment methods, Ahmed et al. (2020a) published a critical overview of different advanced degradation methods—advanced oxidation, PFAS defluorination, advanced reduction, and thermal and nonthermal processes—for PFAS removal from water and wastewater. However, the authors noted that most of these methods are laboratory studies at this stage that show promise but have not been tested in real commercial-scale plants.Fluorine, the lightest halogen, is the most electronegative element; as such, the C–F bond is among the strongest of known covalent bonds (Kissa 2001). Properties of PFAS are useful for industrial and commercial purposes, and mass production has occurred since the early 1950s (Kissa 2001). The same properties, such as thermal stability and chemical inertness, that make them useful also make them stable and environmentally persistent. As production continued into this century, environmental accumulation of PFAS from intentional and nonintentional releases has occurred. For instance, PFAS utilized in specific firefighting foams are distinct sources of environmental PFAS (Dauchy 2019). More broadly, the USEPA states that manufacturing plants, fire-training areas, and fire-suppression activities at airports, refineries, and military installations are primary sources of PFAS release to the environment (USEPA 2021a).The USEPA maintains an online, continuously growing (currently >9,000 chemicals) master list of PFAS (USEPA 2021b). Multiple categories of perfluoroalkyl (fully fluorinated carbon atoms) and polyfluoroalkyl (partially fluorinated carbon atoms) compounds exist. Each PFAS has a linear or branched alkyl chain (Kissa 2001) and a perfluoroalkyl group (CnF2  n+1) (Buck et al. 2011). Detailed definitions of PFAS-related terminology are available in Buck et al. (2011), Dauchy (2019), Pancras et al. (2016), and Wang et al. (2017). PFAS diversity is tabulated in Winchell et al. (2021a), and representative chemical structures are illustrated in Winchell et al. (2021b). PFAS comprise a large family of chemicals cataloged across various nomenclatures such as (1) polymeric (including fluoropolymers, perfluoropolyethers, and side-chain fluorinated polymers) and nonpolymeric [including perfluoroalkyl acids and per-/polyfluoroalkylether acids (PFAA), PFAA precursors, and others], most of which are further subcategorized; (2) legacy, transformation products, and emerging PFAS; (3) ultrashort-, short-, and long-chain PFAS; and (4) polar/nonpolar, nonvolatile/semivolatile/volatile, and neutral/anionic/cationic/zwitterionic PFAS. PFAA precursors are fluorinated chemicals that can be transformed abiotically or biotically into terminal perfluoroalkyl carboxylic acid (PFCA) or perfluoroalkyl sulfonic acid (PFSA) products (Casson and Chiang 2018).In 2006, the USEPA initiated the 2010/2015 PFOA Stewardship Program, inviting eight major PFAS industries to work toward the elimination of perfluorooctanoic acid (PFOA), PFOA precursor chemicals, and related higher homolog chemicals. The program is one of the main drivers of the reformulation of PFAS-based products. All companies have met the PFOA Stewardship Program goals (USEPA 2021c). The Stockholm Convention has restricted the production of perfluorooctanesulfonic acid (PFOS) (UNEP 2019a) and banned the production of PFOA (UNEP 2019b). However, as will be confirmed in this review, these PFAS still occur in the water cycle because of their persistence and because importation from other global markets remains a route for material entering the domestic chain of commerce.In response to increased regulation of long-chain so-called legacy PFAS [eight or more carbons (C8) in the alkyl chain and no ether bonds], industrial applications have shifted toward short-chain (C4–C7) and ultrashort-chain (C2 and C3) PFAS alternatives, with or without ether bonds (Wang et al. 2019; Mulabagal et al. 2018). As reported by the Interstate Technology and Regulatory Council (ITRC 2020), alternative PFAS-based chemical replacements are being marketed and some are appearing in the environment (Munoz et al. 2019; Gustavsson et al. 2018; Hopkins et al. 2018; Wang et al. 2015b), especially several subclasses of ether-PFASs, including the most well-known F-53B [major component (9-chlorohexadecafluoro-3-oxanonane-1-sulfonic acid) or 9Cl-PF3ONS and minor component (11-chloroeicosafluoro-3-oxaundecane-1-sulfonic acid) or 11Cl-PF3OUdS], Gen-X (hexafluoropropylene oxide dimer acid or HFPO-DA), and ADONA (4,8-dioxa-3H-perfluorononanoic acid) compounds (named in their acid forms). These PFAS replacement chemicals are now included in the EPA Analytical methods 533 and 537.1 for PFAS in drinking water (USEPA 2019). Unfortunately, many recently adopted fluorinated alternatives and precursors are more persistent and more environmentally mobile than legacy PFAS (Ghisi et al. 2019).Broad environmental PFOS contamination has been documented. Giesy and Kannan (2001) were the first to report global PFOS contamination in wildlife. For example, PFOS concentrations in the Midwestern US was as high as 2,570  ng/mL in the blood plasma of bald eagles and 3,680  ng/g wet weight in liver samples from mink. Animals tested from urban areas yielded higher concentrations than more remote locations; even wild polar bears contained detectable levels of PFOS. PFOS was measured at higher concentrations in predator species than prey, suggesting bioaccumulation of PFOS. A follow-up study demonstrated bioaccumulation in aquatic species (Giesy et al. 2010) and estimated water quality criteria for the protection of aquatic organisms and wildlife.Several studies across the globe have recently been published to illustrate the broad environmental contamination of PFAS. Muir et al. (2019) documented the wide presence of PFASs throughout the Arctic Ocean and near-shore environments, but they noted that “most long-term time series show a decline from higher concentrations in the early 2000s.” Ali et al. (2021) measured Saudi Arabian Red Sea PFAS concentrations in sediments and edible fish and pointed to wastewater effluents as the main source of these compounds. Bai and Son (2021) performed a study investigating the presence of 17 specific PFAS compounds in 6 surface water and sediment locations in the US state of Nevada, and found that short-chain PFAS were more prevalent in the aqueous phase, while long-chain PFAS were more prevalent in soil (sediments). Cao et al. (2019) investigated soil, sediment, and aqueous-phase PFAS in the Yuqiao reservoir in Tianjin, China, while Chen et al. (2021) investigated similar environmental samples in the Pearl River in southern China. Additional local studies across the globe regarding PFAS environmental contamination include Florida (Cui et al. 2020), Kampala, Uganda (Dalahmeh et al. 2018), Three Gorges Reservoir, China (Jin et al. 2020), the Asan Lake region of South Korea (Lee et al. 2020), South African estuaries (Olisah et al. 2021), Beibu Gulf, South China (Pan et al. 2021), Gangetic Plain, Patna, India (Richards et al. 2021), Mexico City (Rodríguez-Varela et al. 2021), Xiamen Bay, China (Wang et al. 2020), and the Loess Plateau, China (Zhou et al. 2021).In soils, Brusseau et al. (2020) compiled data from more than 30,000 samples from over 2,500 sites globally; the maximum reported PFOS concentrations reached up to several hundred milligrams per kilogram, even in remote regions, where sites were far from PFOS sources. Zacs and Bartkevics (2016) measured PFOA and PFOS in surface water, wastewater, biota, sediments, and sewage sludge in the Baltic region. They noted a prevalence of PFOS over PFOA in surface water and biota samples, whereas mean concentrations of PFOA were greater than PFOS in wastewater, sediments, and sewage sludge. Other, recent, reviews focused on fate and transport within soils and sediments, such as Willemsen and Bourg (2021), who studied the adsorption kinetics of various PFAS structures. Ahmed et al. (2020b) noted decreasing adsorption onto activated carbon as the carbon chain length increased in PFAS, although longer-chain PFAS have larger partition coefficient values than shorter-chained versions. Using numerical studies of PFOS transport within unsaturated soil, researchers have concluded that groundwater contamination could occur from decades to centuries after an initial surface spill (Mahinroosta et al. 2021; Sánchez-Soberón et al. 2020).A growing body of literature has demonstrated that PFAS are found in nearly all environments and in the organisms living therein, including humans. There are now thousands of research reports on the occurrence of PFAS in humans, with data being reported on PFAS levels in human blood, urine, milk, and other tissues (hair and nails) (De Silva et al. 2021; Liu et al. 2020; Jian et al. 2018). While most studies, to date, have focused on human blood, more recent research indicates the distribution of PFAS in various tissues is a function of both PFAS chain length and physiological characteristics (Jian et al. 2018).Water and diet are major potential pathways of PFAS exposure in humans along with food packaging, cookware, air, and air-suspended dust (Ghisi et al. 2019). Elevated PFAS water contamination from localized sources can impact raw water, and thus consumption of contaminated drinking water is a major pathway for human exposure (Tröger et al. 2021; Appleman et al. 2014) if PFAS-specific drinking water treatment is not implemented. The USEPA has established the drinking water health advisory level (HAL) at 70 parts per trillion (combined concentrations of PFOA and PFOS) (USEPA 2016). If drinking water meets the USEPA HAL, the daily intake of PFOS and PFOA from water is less than 20% of the estimated total average human intake per day (USEPA 2016).PFAS contamination of food may occur via irrigation and biosolid application in agriculture (Ghisi et al. 2019) and home food production (Huset and Barry 2018) or farmed and hunted animals (Death et al. 2021), resulting in additional human exposure. Ramakrishnan et al. (2021) documented the presence of PFAS even within organic farming systems and produce production. They noted the use of PFAS-containing composts and feeds, and report that, in some instances, organic meats may contain even higher concentrations of select chemicals than conventionally grown produce. With ubiquitous contamination and bioaccumulation tendencies, growing public concern about human exposure to PFAS stems from an increasing body of evidence of adverse toxicological effects (Sunderland et al. 2019; ATSDR 2021).Given the concern and broad environmental impact of PFAS, this manuscript aims to establish the connections within and between water resource recovery facilities (WRRFs) and drinking water treatment plants (WTPs) including their role in the fate and transport of PFAS cycling in the environment in a manner to inform the industry on this issue. The term water resource recovery facility is adopted here instead of wastewater treatment plant (WWTP) to more broadly reflect the ability to recover valuable resources from wastewater as advocated by the Water Environment Federation (WEF) (WEF 2014). 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