A Critical Review on PFAS Removal from Water: Removal Mechanism and Future Challenges

Abstract

Per and polyfluoroalkyl substances (PFAS) have been extensively employed in a broad range of manufacturing and consumer goods due to their highly persistent nature. PFAS exposure is recognized to pose serious health hazards; therefore, addressing PFAS pollution in water has become a top priority for public health and environmental protection organizations. This review article focuses on the efficiency of different removal techniques (activated carbon, biochar, ion exchange resin, membrane filtration, reverse osmosis, metal-organic frameworks, foam fractionation, ozone fractionation, and destruction techniques) for eliminating different types of short- and long-chain PFAS from water. Hydrophobicity and electrostatic interactions are revealed to be the primary mechanisms for the elimination of PFAS. The efficiency of all techniques to eradicate short-chain PFAS is comparatively lower compared to long-chain PFAS. The destruction techniques are the most efficient but have some drawbacks, including the formation of PFAS precursors and high operational costs. According to the findings from the study, it is anticipated that combined methods will be required to effectively remediate PFAS-contaminated water.

1. Introduction

Per- and polyfluoroalkyl substances (PFAS) are man-made fluorine-containing organic compounds exhibiting distinctive physiochemical characteristics due to the strong carbon and fluorine bonding (C-F) [1,2]. PFAS are comprised of hydrophobic and oleophobic carbon chains in which the hydrogen can be completely or partly substituted with fluorine atoms [3,4]. Perfluoro carboxylic acids (PFCAs) and perfluoroalkane sulfonic acids (PFSAs) are the two well-studied major subclasses of PFAS [5]. They are also recognized as “forever chemicals” because of their persistent nature. Over 5000 different forms of PFAS exist in the world [6]. Due to the exceptional resistance properties of PFAS to heat, oil, and water, they are extensively used in different areas, including domestic (nonstick cookware), firefighting foams, and industrial food packaging applications [7]. It has been reported that PFAS concentrations in aqueous medium vary from pg/L to mg/L [8]. According to Hopkins et al. [9], the Cape Fear River, a water supply utilized by utilities in North Carolina, possesses high quantities of different PFAS [9]. In September 2015, PFAS contamination was identified by the New South Wales Government of Australia in Williamtown. Subsequently, in August 2016, Australia’s Department of Defense identified the key exposure route of PFAS contamination, causing health hazards by contaminating drinking groundwater [10].

Six illnesses have been linked to PFOA exposure in a study conducted involving more than 68,000 people living near a manufacturing factory in the United States, including testicular cancer, high cholesterol, pregnancy-hypertension, kidney cancer, thyroid disease, and ulcerative colitis [11]. PFAS have also been found in umbilical cord blood, breast milk, and human blood [12,13]. COVID-19 is known to be worsened by pulmonary buildup of PFAS, and the latest evidence found that increased plasma-PFAS concentrations in human blood are linked to this severity [14]. Over 95% of Americans have PFAS in their blood [5,15,16]. Human exposure to PFAS could happen through PFAS-contaminated food consumption (grown on PFAS-contaminated soil) [17]. PFAS may infiltrate soil through a variety of routes. PFAS concentration in soil ranges from nanograms to micrograms [18]. From the soil, the PFAS could leach down to contaminate the groundwater, which is another route of human exposure to PFAS [19]. The Stockholm Convention on Persistent Organic Pollutants (POPs) placed PFOA on its Annex A (elimination) list and PFOS on its Annex B (restricted) list due to the prolonged toxicity and biomagnification of both substances for a worldwide ban [20]. The Environmental Protection Agency (EPA) forbids corporations in the United States from producing, exporting, or importing items that include specific long-chain PFAS without first undergoing EPA inspection and clearance [21].

Short-chain PFAS (C2–C7) have been used as an alternative to PFOS and PFOA (C8) due to their decreased production [22]. In 2009, Dupont developed the PFOA alternative GenX (C6), also known as hexafluoropropylene oxide dimer acid (HFPO-DA), which was later commercialized by Chemours, a DuPont subsidiary. Due to its widespread use in industry as a less harmful replacement, GenX has gained greater prominence compared to other short-chain PFAS [22]. It has been detected in river water samples collected from different countries, including the United States, China, Germany, and the Netherlands. The GenX concentration in those samples was higher than the permissible limits of PFAS and PFOS set by the US EPA [23]. In recent investigations, GenX has been shown to be more hazardous and bioaccumulative compared to PFOA, and the efficiency of the drinking water plant to remove GenX was quite poor [24,25,26]. PFAS are very difficult to remove from the environment due to their persistent physiochemical and thermostable nature; however, owing to their increased solubility and mobility, the shorter ones are more challenging to eliminate [2,8]. Numerous techniques have been used to remove the PFAS and GenX from the water, including electrooxidation [24], activated carbon (AC) [27], membrane technologies [11,28], electrochemical degradation, clay minerals [8], ion-exchange technologies [7], nanoparticles [29,30], reverse osmosis [31], biochar [32,33], etc. However, research studies on GenX removal are still scarce. The efficacy of various methods for eradicating GenX and PFAS must be compared. Long- and short-chain PFAS, such as GenX, are the main focus of this assessment, along with their background, removal methods, removal processes, cost analysis, and comparison of efficiency across various technologies.

History and Sources of PFAS

Roy J. Plunkett accidentally discovered a fluoropolymer known as polytetrafluoroethylene (PTFE) on 6 April 1938, while doing an experimental study on DuPont Freon refrigerant gases. Ten years of study culminated in 1941 with the patenting and commercialization of PTFE under the brand name Teflon [34]. In 1947, another company named 3M invented its trademark PFOA with the invention of numerous chemical compounds having eight carbons tied to fluorine [35]. In 1963, scientists in the US Navy worked with 3M to manufacture aqueous film-forming foams and patented them in 1967 [36]. After that, different forms of PFAS were developed by numerous industries, e.g., packaging, paint, cleaning, and fabric industries, etc. A West Virginia farmer named Wilbur Tennant filed the first complaint against DuPont in 1999, accusing the company of releasing PFAS into local water supplies. Tennant saw several health problems, including black teeth, birth deformities, and cancers in his cattle. More than a hundred of his animals died after consuming water from a nearby creek that had been poisoned by PFAS sludge discharged by DuPont in the 1980s [35]. A summary of PFAS events history is displayed in Figure 1.

After that, PFAS were found to cause cancer, immune system suppression, disturbance of thyroid hormones, and low birth weight in infants [37,38]. As a result, efforts have been executed to mitigate PFAS environmental effects. The two most produced and utilized PFAS are PFOA and PFOS. Both are highly persistent and bioaccumulative [39,40]. Thus, the Stockholm Convention’s POPs list restricted their use. The PFOA Stewardship Program, launched by the US Environmental Protection Agency in January 2006, has accelerated the process of phasing out this chemical [41]. Despite having similar harmfulness and persistence, the short-chain substituents are thought to be less bioaccumulative and so represent less risk to human health. The chemical stability of the PFOA alternative GenX is low because of the presence of etheric bonds between the perfluorinated moieties [42]. According to the studies, GenX levels in the water systems were greater than the combined EPA recommendations for PFOS and PFOA. GenX, as shown in recent research, poses a significant challenge to water treatment facilities because of its high toxicity and bioaccumulation, which are similar to or superior to PFOA [24,25].

Table 1. Physiochemical characteristics of PFAS reported in this study.

Type	Abbreviation	Molecular Formula	IUPAC Name	Molecular Weight	Structure
PFECAs	Per- or polyfluoroalkyl ether carboxylic acid
PFMOPrA	Perfluoro-2-methoxypropanoic acid	C4HF7O3	Perfluoro-2-(perfluoromethoxy)propanoic acid, 2,3,3,3-Tetrafluoro-2-(trifluoromethoxy)propanoic acid	230.04
PFO2HxA	Perfluoro(3,5-dioxahexanoic) acid	C4HF7O4	Perfluoro-3,5-dioxahexanoic acid	246.04
PFMOBA	Perfluoro-4-methoxybutanoic acid	C5HF9O3	Perfluoro-4-methoxybutanoic acid, 2,2,3,3,4,4-Hexafluoro-4-(trifluoromethoxy)butanoic acid	280.04
HFPO-DA	Hexafluoropropylene oxide dimer acid	C6HF11O3	Perfluoro-2-methyl-3-oxahexanoic acid	330.05
PFSAs	Perfluorosulfonic Acid
PFBS	perfluorobutane sulfonate	C4HF9O3S	1,1,2,2,3,3,4,4,4-Nonafluorobutane-1-sulfonic acid, Nonafluorobutanesulfonic acid	300.1
PFPeS	Perfluoropentanesulfonic acid	C5HF11O3S	perfluoropentanesulfonic acid, 1,1,2,2,3,3,4,4,5,5,5-undecafluoropentane-1-sulfonic acid	350.11
PFHxS	Perfluorohexanesulfonic acid	C6HF13O3S	Perfluorohexane-1-sulphonic acid, 1,1,2,2,3,3,4,4,5,5,6,6,6-Tridecafluorohexane-1-sulfonic acid	400.12
PFHpS	Perfluoroheptanesulfonic acid	C7HF15O3S	perfluoroheptane sulfonic acid, 1,1,2,2,3,3,4,4,5,5,6,6,7,7,7-pentadecafluoroheptane-1-sulfonic acid	450.12
PFOS	Perfluorooctanesulfonic acid	C8HF17O3S	Heptadecafluorooctanesulfonic acid, 1,1,2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-Heptadecafluorooctane-1-sulfonic acid.	500.13
PFNS	Perfluorononanesulfonic acid	C9HF19O3S	Nonadecafluoro-1-nonanesulfonic acid, 1-Nonanesulfonic acid, 1,1,2,2,3,3,4,4,5,5,6,6,7,7,8,8,9,9,9-nonadecafluoro-	550.14
PFDS	Perfluorodecanesulfonic acid	C10HF21O3S	Henicosafluorodecanesulphonic acid, 1,1,2,2,3,3,4,4,5,5,6,6,7,7,8,8,9,9,10,10,10-henicosafluorodecane-1-sulfonic acid	600.15
PFCAs	Perfluoroalkyl carboxylic acids
PFBA	Perfluorobutanoic acid	C3F7COOH	Heptafluorobutanoic acid	214.04
PFPeA	Perfluoropentanoic acid	C4F9COOH	Perfluorovaleric acid, nonafluoropentanoic acid	264.05
PFHxA	Perfluorohexanoic acid	C5F11COOH	Undecafluorohexanoic acid	314.05
PFHpA	Perfluoroheptanoic acid	C6F13COOH	Tridecafluoroheptanoic acid	364.06
PFOA	Perfluorooctanoic acid	C8HF15O2	Pentadecafluorooctanoic acid, 2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-Pentadecafluorooctanoic acid	414.07
PFNA	Perfluorononanoic acid	C9F17O2	heptadecafluorononanoic acid, Perfluoro-n-nonanoic acid	464.08
PFDA	Perfluorodecanoic acid	C9F19COOH	Nonadecafluorodecanoic acid, Perfluoro-N-decanoic acid	514.08

demonstrates some physiochemical characteristics of the different types of PFAS reported in this study with their abbreviation given below:

There are two types of PFAS emission sources: direct emission sources and indirect emission sources. The direct emission sources include emissions at any time in the lifecycle when products contain the substance as a primary or residual impurity. The indirect emission sources refer to the formation of the product from the transformation of the product’s precursors in the environment. Primary PFAS producers, AFFF users, and downstream sectors (such as leather, metal plating, semiconductors, and textiles) are all direct contributors to PFAS contamination [43]. The two main indirect sources of PFAS contamination are water treatment plants and municipal landfills [43]. Some common sources of PFAS are exhibited in Figure 2.

AFFF (Airforce aqueous film-forming foam) formulations often include PFAS due to their performance as surfactants for Class B hazardous fuel fires [44]. The AFFF produced by 3M using electrochemical fluorination (ECF) technology is known as legacy PFOS AFFF [45]. The military, aviation industry, oil and gas industries, and the firefighting sector were among the first to use AFFF due to their superior efficacy and efficiency in combating fires [45,46].

Flame retardants in electronics may be either brominated, like polybrominated diphenyl ethers (PBDEs), or fluorinated, like PFAS [47]. As a flame retardant, polycarbonate resins often include the potassium salt of perfluorobutane sulfonate (PFBS). Lithium salts of PFAS are used as electrolytes in fuel cells and batteries [48]. Each year, we produce more than 50 million metric tons of e-waste, which includes all forms of electronic and electrical equipment as well as their components that have been discarded [49]. Electronic waste processing plants and landfills that accept abandoned electrical and electronic equipment discharge toxic substances, such as PFAS, into the environment. The leachate that is produced during landfill treatment and then treated locally or transported to wastewater treatment facilities contains PFAS compounds, and there is currently no efficient way to remediate or remove them [48,50].

New PFAS compounds, including fluorotelomers, require further study in wastewater because of their widespread use in industries like metal plating, oil production, fluorinated impregnations, and food packaging [51]. There is evidence from studies indicating shorter-chain PFAS compounds have increased in prevalence after the prohibition of longer-chain PFAS compounds [52]. Wastewater treatment plants (WWTPs) are used to treat wastewater from a variety of sources, including but not limited to landfill leachates, agricultural wastewater, and effluents produced from home and industrial sources. Even though many pollutants are degraded throughout the WWTP process, PFAS are resistant to the typical treatment procedures used in WWTPs [53]. Due to the transformation of large-chain precursor PFAS compounds into smaller-chain PFOS during treatment, PFAS concentrations in the effluent are frequently higher than those in the influent [51,54]. PFAS are less known for its use in cosmetics and other personal care items. The perfluoroalkyl moiety gives compounds interesting characteristics such as strong surface activity, resistance to water and oil, and chemical and thermal stability [43]. U.S. Environmental Protection Agency researchers have shown that PFAS may be found in a variety of common home items, including food packaging materials [55]. To improve their water- and oil-resistance, several types of packaging include a PFAS coating. In the end, these items are often thrown in landfills where they seep into different bodies of water.

2. Treatment Techniques for the Removal of PFAS

At present, water utilities have few viable methods for PFAS removal. Traditional filtration methods like GAC filters, reverse osmosis, and ion-exchange resins have proven to be effective in removing PFAS but they were not developed to precisely bind these emerging pollutants. The PFAS molecules might pass through these technologies without destruction or sorption. Researchers are currently building PFAS-specific adsorbents and treatment procedures for the complete destruction of PFAS molecules rather than just capturing and storing them [56]. The following are the recent technologies used by researchers for the treatment of PFAS-contaminated wastewater.

2.1. Sorption

2.1.1. Biochar

Biochar is a carbonaceous, highly stable, porous, and environment-friendly material. Biochar is primarily made from the abundant, low-cost, and widely available biowastes of the agricultural and food industries [57]. The surface of biochar is comprised of different types of functional groups, mainly heterocyclic and aromatic, including hydroxyl, carboxyl, carbonyl, etc. which assist in PFAS adsorption [58]. Hydrophobicity and electrostatic interactions are the main mechanisms of PFAS removal using biochar [59]. Biochar has been shown to be an effective method for detoxifying water of pharmaceuticals and heavy metals [60,61,62]. In addition, it is possible to tailor its exceptional physical features (like pore size distribution) to the specific needs of a variety of applications (like the removal of short-chain PFAS).

In order to remove PFAS from wastewater, Patel et al. produced a unique economical biochar for PFAS removal from the byproduct of anaerobic digestion of organic contaminants in the wastewater treatment process [59]. The anaerobic digestion of organic pollutants releases biogas, which can be converted into hydrogen and carbon nanomaterials (CNMs) by sequestering the carbon during catalytic decomposition. The carbon nanomaterials were loaded on the biochar (CNM-s loaded biochar). The CNMs-loaded biochar exhibited a higher removal efficiency (79%) for PFAS compared to ilmenite-modified biochar (54%). Softwood and hardwood shavings were used to prepare biochar by pyrolyzing at varied temperatures (300, 400, and 600 °C) for PFAS removal. These biochars were post-modified by thermal oxidation at 400 °C for half an hour which results in a high surface area and increased pore size. Once PFAS has been adsorbed on biochar, they are destroyed by heat activation at 500 °C in the presence of air or nitrogen. The biochar shows higher performance for PFAS removal in the next cycle after the thermal activation [63]. In another study conducted by Wang et al., softwood and hardwood shavings were used to fabricate the biochar at different pyrolysis temperatures of 300–600 °C [64]. After pyrolysis, the material is subjected to hot air oxidation, which boosts the material’s ability to absorb by increasing the surface area and pore size. This process improves the reactivity of biochar made from softwood for some PFAS by threefolds compared to untreated biochar. After being exposed to a solution containing PFCAs, PFSAs, and GenX, the materials were set aside to dry, followed by heating the dried materials at 500 °C in the air for 30 min to eliminate the PFAS. After being treated, the biochars and granular activated carbon (GAC) were put through further sorption tests, this time with a combination of PFBA, PFPeA, GenX, PFHpA, and PFOA. Reactivated biochar virtually always had a higher adsorption-desorption distribution coefficient (Kd) for all PFAS than raw charcoal.

Krahn et al. used an effective waste management practice for the fabrication of biochar by pyrolyzing it at 700 °C using sewage sludge (SSBCs) and wood chips (WCBC) waste for the elimination of short-chain per fluorinated carboxylic acids (PFPeA, PFHxA, and PFHpA) and long-chains (PFOA, PFNA, and PFDA) and the adsorption of all except PFPeA and PFHxA onto two biochar’s (SSBC and WCBC) followed a strong log-linear relationship (r² > 0.9) [65]. These poor fits may be explained by the fact that ionic interactions are more crucial for the sorption of these short-chain PFCAs than hydrophobic ones are for the sorption of the longer-chain PFCAs.

Two-step pyrolysis or post-modification was utilized for producing the biochar in other research that used discarded coffee grounds. Two different approaches were investigated for functionalizing the biochar surface with nitrogen-containing groups on the surface, which would then allow for radical-initiated polymerization. The first approach was electrophilic aromatic substitution, carried out by mixing exact amounts of pristine biochar with nitric and sulfuric acid, following the reduction of the nitro group to -NHx using acetic acid, ammonium hydroxide, and sodium dithionite. In the second approach, they added melamine (BC-M) or ammonium chloride (BC-N) to the biochar water mixture and then heated it to initiate the nitrogen modification. To improve the adsorbent’s selectivity for PFAS, a layer of molecularly imprinted polymers (MIP) was added onto the surface of the nitrogen-modified biochar using thermally stimulated radical-initiated polymerization. The BC-M-MIP exhibited higher Kselectivity for PFBS (4.52) and PFOA (3.76) compared to unmodified biochar at 0.043 and 0.039 mg PFAA/g*g/m², respectively [66]. In another experiment, tree bark and eucalyptus biochar were used for the removal of different chain lengths of PFAS (4 to 11 fluorinated carbons). The sorption increased with the chain length, which points out hydrophobic interactions as an important mechanism in the adsorption process [67]. Wu et al. used different feedstocks (switchgrass (SG), water oak leaves (WO), and biosolids (BS)) for the fabrication of biochar and modified them using FeCl₃ and carbon nanotubes (CNTs), which increased their surface area and porosity [68]. The fastest adsorption equilibrium was achieved by the oak leaf-biochar composite. The biosolid-biochars in either pristine or modified form exhibited higher adsorption for PFOA compared to the other two biochars. The maximum adsorption capacity for PFOA was exhibited by BS-Fe (469.65 μmol/g), and the minimum adsorption potential of 39.54 μmol/g was exhibited by WO-CNT (Figure 3).

Gasification at 900–1000 °C was used to create biochar from Douglas fir in another investigation, using syngas as a byproduct. To modify the biochar with Fe₂O₃, iron (III) chloride (18 g), and iron (II) sulfate (36.6 g) were added to a solution (2 L) comprising BC (50 g). Next, after keeping the pH constant at 10 for 24 h while gently introducing 10 M NaOH, Fe₃O₄ particles were produced and deposited onto BC. The Fe₂O₃ modified biochar was ready after vacuum filtration, washing (3x ethanol), and drying (50 °C) step. The Fe₂O₃-modified biochar removed 75–85% of PFBS, GenX, and perfluorooctane sulfonamide (PFOSA) compared to PFOS (82%). The primary mechanisms for their adsorption onto Douglas fir Fe₂O₃-modified biochar were determined to be electrostatic and hydrophobicity interactions [69]. Modified or activated biochar exhibited higher adsorption potentials for both short- and long-chain PFAS compared to pristine biochar. The primary mechanism of PFAS removal by biochar hydrophobic and electrostatic attraction. The removal of short-chain PFAS is comparatively low; therefore, further investigation on sustainable removal methods specifically targeting short-chain PFAS during wastewater treatment is needed.

2.1.2. Activated Carbon

Granular activated carbon (GAC) adsorption has been shown to be an effective strategy for detoxifying water contaminated with long-chain PFAS. According to the Web of Science database, activated carbon (AC) is at the top of the list of adsorbents (40% of PFAS studies) for PFAS removal. The non-polar and hydrophobic surface of AC significantly interacts with hydrophobic PFAS, leading to its extensive usage in adsorption processes [70,71]. In recent years, GAC has been utilized in several trial experiments to investigate the impact of PFAS features, such as isomer structure (branched or linear), perfluorocarbon chain length, and functional groups (carboxylates or sulfonates), on the elimination of different PFAS. On aggregate, GAC was shown to be around 80% and 90% efficient in removing PFAS [72,73,74]. Activated carbon has become the most widely utilized sorbent for PFAS removal because it may function both as a single intervention and as part of a pre-treatment step.

Mohamed et al. used a cost-effective method to fabricate AC using sludge-based activated carbon (SBAC). Domestic wastewater treatment facility sludge was used in this study [75]. The manufactured sewage sludge (SS) was dried and pyrolyzed (at 500 °C for 2 h), then treated with 2.5 M ZnCl₂ with agitation. The optimal SBAC was developed at activation time (0.5, 0.75, 1, 1.5, and 2 h), ZnCl₂ concentration (1.5, 2, and 2.5 M), and pyrolysis temperature (400, 450, and 500 °C). To evaluate the viability of SBAC manufacturing, a life cycle cost analysis (LCCA) was also carried out. Most altered SBACs with a ZnCl₂ impregnation ratio of 2.5 M removed PFAS and MB effectively (95.1–100%), except for the sample generated at a low temperature (400 °C). For short-chain PFAS chemicals, this sample had the least adsorption rate (35.3–57.8%). The high solubility and poor hydrophobicity of the SBAC made at 400 °C may explain why it has a low short-chain PFAS sorption potential. The optimal SBACs generated at 1.5 and 2 M ZnCl₂ under varied pyrolysis durations were examined. Consequently, PFAS and MB were removed at 90.6–100%. The reduction of the ZnCl₂ ratio from 2.5 to 1.5 reduced the possibility of mineral depletion and ecotoxicity of terrestrial, marine, and freshwater by 38%, 24%, 22%, and 49%, which are the most impacted factors. By comparison to natural adsorbents like kaolinite, whose surface area is measured in the tens of m²/g, AC’s is 100 m²/g. Furthermore, commercial AC is notably more costly than natural adsorbents (>4 USD per kilogram) [75,76]. Research by Mohammad et al. found that the price of producing SBAC may be brought down to 1.2 USD/kg [75]. After adsorbing short-chain PFAS, AC must be regenerated, which substantially increases the PFAS treatment price [77]. Meanwhile, Zhong et al. found that the biologically activated carbon (BAC) with different age periods of 1, 4, 7, and 13 years were analyzed for the sorption of 14 PFAS types from drinking water treatment plants (DWTPs) [78]. As a result of their surface and microporous nature, bacteria may cling to GAC and transform it into BAC via the formation of biofilms. Neither the carbon maturity nor the depth had any effect on the microbial species present in the BAC biofilm; rather, the major difference was population. Bacteria of the proteobacterial family have a significant impact on BAC’s ability to remove PFAS. Raw water had a total PFAS content of 127.4 ng/L. The total PFAS concentration decreased to 101.9 ng/L following drinking water treatment plant remediation. In terms of the relative effectiveness of the various methods of removal, pre-ozone treatment led to a rise in PFAS content, from 127.4 ng/L to 142.7 ng/L. The occurrence of its intermediates changed to PFHxA by ozonation likely accounts for most of the rise in PFAS content [79]. Treatment with ozone biological activated carbon (O3-BAC) was found to be promising in removing PFAS from water (20.74%), indicating that O3-BAC plays a pivotal role in the treatment of water to eliminate PFAS [78]. Riegel et al. conducted lab and field experiments to examine the adsorption potential of five different GACs (lignite-based GAC, Coconut shell GAC, bituminous coal-based GAC, and some unknown raw material) for short- and long-chain PFAS removal from drinking water [80]. In the laboratory experiment, demineralized water (contaminated with four PFAS having an initial concentration of 6 μg/L: PFOA, PFBA, PFHxA, and PFPeA) was used, whereas PFAS-contaminated groundwater was used for the field column trial in the Southern Germany waterworks, which was contaminated due to paper sludge. As the polarity of PFAS decreases with increasing chain length, GAC is more effective in adsorbing long-chain PFAS versus short-chain PFAS [81]. The adsorption of PFSA was higher compared to PFCA [82]. The results of the AC sorption experiments indicated that PFOA was the most readily adsorbed, followed by PFHxA, PFPeA, and finally PFBA. In the field column experiment, the long-chain PFOA, as well as the short-chain PFHpA and PFHxA, were both efficiently removed by GAC treatment. Nevertheless, adsorption by GAC is either ineffective for PFBA and PFPeA or leads to short operating durations of GAC adsorbers, causing expensive treatment costs [80]. In a water utility, Burkhardt et al. performed pilot research consisting of two parts: in the first, nine PFAS compounds were investigated; in the second, sixteen PFAS compounds were analyzed; and in both parts, five different GACs were used and pore and surface diffusion models were used for the prediction of outcomes [27]. The GAC treatment was shown to be effective in removing all the PFAS tested in this investigation. Nevertheless, the remediation effectiveness differed significantly among the two pilot phases, which were carried out at various periods of the year, demonstrating that the PFAS had considerably varying adsorption properties that also fluctuated periodically. Granular activated carbon (GAC) was shown to be an effective remediation approach for PFAS removal in this pilot study, meeting the treatment objectives of the utility that conducted the research [27]. Mini-column studies at the research lab level were used to assess the effectiveness of GAC in removing PFOA and PFOS, two typical persistent PFAS, from drinkable water at three separate treatment facilities in Ontario, Canada [83]. Column studies were performed utilizing both sterilized and unsterilized GAC to differentiate between the adsorption and the possible degradability of PFOA and PFOS over the GAC. GAC service duration, temperature, influent concentration, and empty bed contact time (EBCT) were studied for their impact on PFOA and PFOS removal. The GAC samples were found to eradicate 40–55% of PFOS and 20–40% of PFOA, apart from Plant B, where removal of PFOA was insignificant due to 9-year-old GAC, which might be due to natural fouling of organic matter (OM) on the GAC surface. These outcomes are of great importance as they exhibited that the old GAC in drinking water treatment plants with continuous operation for taste and smell regulation were able to remove PFAS without any modification or further action [83].

To sum up, activated carbon is an efficient and common approach for treating water contaminated with PFAS (per- and polyfluoroalkyl compounds). Many variables influence how effectively AC filters remove PFAS from water. These include the type of AC employed, the characteristics of PFAS, and the water quality criteria. The possibility of secondary contamination makes it crucial to give careful thought to how AC laden with PFAS will be disposed of. The downsides of AC, such as the necessity to renew the carbon medium and the possibility of PFAS desorption, are outweighed by its benefits.

2.1.3. Metal-Organic Frameworks

Metal-organic frameworks (MOFs) are crystalline materials with a porous structure made up of inorganic nodes and organic linkers that self-assemble into three-dimensional lattices via coordination bonding [84]. These materials attractive properties include their amenability to modification at multiple levels (nodes and linkers), their tendency toward crystalline structure, and the existence of empty spaces, which have all led to numerous applications that follow the concepts of sorption processes, such as catalysis, separation, and filtration. The elimination of PFAS from water is a promising new use for MOFs [85]. The pore volume and surface area of MOFs are very high, up to 3.9 cm³/g and up to 7130 m²/g, respectively. MOFs are promising options for adsorption-based applications requiring rapid diffusion and kinetics due to their adaptable pore diameters (pore sizes up to 10 nm) and functionalities [86,87,88]. Further study is warranted to enhance the stability of this class of adsorbents because of the low water stability of MOFs, which might be a concern when employing these materials for water treatment. Zeolitic imidazolate framework (ZIF) and chromium (III) terephthalate metal-organic framework (MIL-101) are two examples of water-stabilized porous MOFs that have recently emerged and are showing tremendous promise for use in water treatment operations [89,90]. The adsorption process with MOFs is highly dependent on their physicochemical and geometric parameters, such as morphology, pore size, interactions, shape, and surface area, all of which are influenced by the MOF synthesis technique [91,92].

Chang et al. fabricated a water-stable MOF called PCN-222 using zirconium tetroxide (7.3 mg) as a metal source and tetrakis (4-carboxyphenyl) porphyrim (4.5 mg) as a ligand [90]. The PCN-22 exhibited a maximum adsorption potential of 2257 mg/g for PFOA removal and fitted well to the Langmuir isotherm, which favors monolayer adsorption. The pH of PCN-22 was maintained around 5 to maintain the positive charge on PCN-222. Further increases in pH cause a decrease in the adsorption potential of PCN-222 for PFAS due to the deprotonation of PFOA. The adsorption of PFOA increased with temperature, suggesting the reaction nature is endothermic. The PFOA were adsorbed on PCN-222 via hydrogen bond formation, electrostatic attraction, and hydrophobic attraction.

In another study, a MOF-808 was fabricated using zirconium chloride for nodes and H3BTC ligands for the removal of perfluorobutanesulfonate (PFBS) by senior-level students in a capstone course. Students were able to swiftly synthesize MOF-808 with a high yield in the chemical lab and characterize it using X-ray diffraction (XRD), Thermogravimetric analysis (TGA), and Diffuse Reflectance Infrared Fourier Transforms Spectroscopy (DRIFTS). Adsorption isotherms for PFBS on MOF-808 samples generated by students range from 2.8(8) to 10⁻⁴ M PFBS. When the dilution factor is included, the adsorbate concentration of PFBS rises to 3.0(8) 10⁻⁴ M [93]. The 2 g of Cr(NO₃)₃.9H₂O, 0.83 g of H2BDC, and 0.15 mL of HF were mixed together in 70 mL of distilled water to create MIL-101 (Cr). The MIL-101 (Cr)@AC was fabricated by mixing 50 mg of AC in the precursor solution. To make MIL-101(Cr)-NH₂, 0.4 g of NaOH, 0.72 g of 2-Amino terephthalic acid, and 1.6 g of Cr(NO₃)₃.9H₂O were dissolved in DI water (30 mL) and mixed for 15 min at 160 °C for 16 h. In FTIR outcomes, the C-H bond at 1104 cm⁻¹, C=O bond at 1650 cm⁻¹, and 1620 cm⁻¹ (C=C bond) displayed greater peak intensity compared to MIL-101 (Cr), indicating contact between ligand and AC, but no new functional group was introduced by AC’s interaction with MIL-101 (Cr). All the adsorbents showed an increase in adsorption from pH 2 to 4, followed by a reduction from pH 4 to 10. A comparison of AC’s adsorption effectiveness across a pH scale of 2–10 revealed no discernible differences. This might be because AC maintains a constant surface charge across the specified pH range. The highest adsorption was exhibited by AC (28.248 mg/g), followed by MIL-101(Cr)@AC (25.706 mg/g) (

Table 2. Removal of PFAS via adsorption on biochar, metal-organic frameworks, and activated carbon.

Removal Technique or Adsorbent	PFAS Type	pH	PZC	Dosage	Surface Area (m2/g)	Initial Conc.	Adsorption Capacity (mg g−1)	Regeneration Method	Removal Rate (%)	Ref
Carbon nanomaterial coated biochar (CNM-Biochar)	PFOS	-	-	1 g	40 ± 2.1	310 µg/L	-	-	95	[59]
	PFOA	-	-	-	-	25 µg/L	-	-	71
	PFHxS	-	-	-	-	140 µg/L	-	-	72
Biochar alginate composite	PFOS	8	4.37	0.5–1.5 g/L	4.7	100 μg L−1	1572 µg g−1	-	99	[62]
	PFBS	-	-	1.0–3.0 g/L	-	-	2895 µg g−1	-	40
SW600-PPAO (Soft wood-post pyrolysis air oxidation)	PFBA	7.4	-	400 mg	832	0.0687–4.63 μM PFAS	1.75	-	-	[64]
	PFPeA	-	-	-	-	-	1.75	-	-
	GenX	-	-	-	-	-	1.45	-	-
	PFDA	-	-	-	-	-	4.76	-	-
	PFBS	-	-	-	-	-	2.6	-	-
	PFHxS	-	-	-	-	-	3.79	-	-
Acid-modified biochar (H-SL)	PFOA	7	-	1 g	3.626	50–100 mg/L	49.02	-	-	[94]
	PFOS	-	-	-	-	-	79.8	-	-
Sewage sludge-activated carbon (ZnCl2-impregnated)	PFBA	-	-	5 mg	723	50 μg/L	-	-	87	[95]
	PFPeA	-	-	-	-	-	-	-	95
	PFHxA	-	-	-	-	-	-	-	96
	PFHpA	-	-	-	-	-	-	-	97
	PFOA	-	-	-	-	-	-	-	100
	PFNA	-	-	-	-	-	-	-	100
	PFBS	-	-	-	-	-	-	-	99
	PFHxS	-	-	-	-	-	-	-	98
	PFOS	-	-	-	-	-	-	-	100
Classical powder-activated carbon	PFOS	4	-	0.004 g	1232.38	2 ppm	28.248	50:50 ethanol and water mixture	93	[91]
Thermal defunctionalized CACF (DeCACF)	PFOA	7	9.3	-	1000	8 (µg/L)	80	methanol extraction	-	[96]
	PFOS						103		-
	PFBA						5.1		-
Commercial PAC AquaNuchar (AquaNC)	PFOA	7	4.75	10 mg	1860	0.5 mg	2.46	-	-	[97]
	PFOS						2.48		-
Commercial-activated carbon felt ACTITEX WK L20 (CACF)	PFOS	7	5.9		1100	-	33.3	-	-	[96]
	PFOA						2.08		-
Commercial-activated carbon felt ACTITEX FC 1501 (FC15)	PFOS	7	7.1	-	1600	-	156	-	-	[98]
	PFOA						35.7		-
Commercial-activated carbon felt ACTITEX FC 1001 (FC10)	PFOS	7	6.5	-	1400	-	108	-	-
	PFOA						10.2
MIL-101 (Cr)@AC MOF	PFOS	4	-	100 ppm	6955	2 ppm	25.706	50:50 ethanol and water mixture	80	[91]
PCN-222 MOF	PFOS	4.1–5.3	7.5	0.003 g	1948	500 mg/L	2257	-	-	[90]
UiO-66	PFOS	5	6.4	0.5 g/L	1423	500 mg/L	1.24 mmol	-	-	[85]
MIL-96-Al modified with hydrolyzed polyacrylamide	PFOA	-	-	1.0 g/L	75	1000 mg/L	340	methanol/water	-	[99]
DUT-5-2	PFOA	3.0	-	5 mg	1840	30 mg/L	98.2	methanol	60.9	[100]
	PFOS	-	-	-	-	-	145.4	-	-
MIL-100-Fe	PFOA	3.3	3.64	50–150 mg	1237	50–1000 mg/L	426.6	-	-	[101]

). The most prominent adsorption interactions for AC and MOFs were hydrophobic-hydrophobic and electrostatic, respectively [91].

Molecular simulations were used to examine the efficacy of various fluorine-functionalization techniques on MOFs for the removal of PFAS from water. The first class consists of fluorinated anion-functionalized MOFs, with AF₆⁻² anions (A = Si or Ti) serving as the connecting ligands. The second group consists of substitutions made using trifluoromethyl or fluorine. Finally, perfluoroalkane graft-functionalized MOFs are the third kind of fluorine-functionalized MOFs in this study. Of all the materials studied in this work, two of them, TIFSIX-1-Cu and Zn(4,4′-bpy)2(SiF6), had the greatest Henry’s Law Coefficients (HLCs) for PFOA. The results of this investigation show that fluorination is an effective method for removing PFAS from water resources because it improves the MOF-PFOA interactions [102]. GenX has lesser adsorption on AC than other legacy PFAS like PFOA (Wang et al., 2019). Commercialization of carbon-based adsorption technologies is hindered by issues with adsorbent regeneration, tuning adsorbents with favorable characteristics, and OM competition. Commercialization of carbon-based adsorption technologies is hindered by issues with OM competition, adsorbent renewal, and the design of adsorbents with the necessary surface chemistry/properties [103,104].

The large surface area, customizable pore widths, and metal coordination sites of MOFs make them a desirable option for removing PFAS from water sources that have been polluted. The short-chain PFAS have received less attention from researchers due to their less absorbable and more persistent nature. To remove PFAS more effectively, further study is required to enhance MOF features, such as enhancing selectivity for particular PFAS molecules and strengthening stability under different environmental conditions.

2.1.4. Ion Exchange Resins

Polymers’ strong adsorption capabilities make them attractive for PFAS elimination. Investigation into PFAS remediation utilizing ion exchange resin (IXR) has risen dramatically in the past few years in consideration of the material’s established adsorption potential for PFAS [105]. Research on PFAS adsorption using anion exchange polymers or resins (AER) has been conducted most often. The adsorption mechanism of PFAS by AER is based on hydrophobic and selective ion-exchange interactions [106]. Chow et al. conducted a pilot study in which they examined the adsorption behaviors of PFAS by GAC and two strong-base gel anion exchange resin (AER) columns that were operating in parallel to clean up contaminated groundwater over the course of 441 days. In terms of mass, GAC was able to hold onto 85 μg PFAS/g adsorbent, with 71% PFSA and 28% PFCA. The AER1 and AER2 adsorbed 441 and 460 μg/g of PFAS. The PFCA adsorption percentage was 11–12%, while PFSA adsorption potential was 88–89% by AER, which is seven times more than PFSA. The amount of PFAS absorbed by AER was more than five times that of GAC. Gel AERs performed slightly better than GAC for long-chain PFCA and significantly better for all PFSA in terms of mass adsorption [107]. In another experiment, Pluronic^® F-127 and Abberlite^® IRA910 beads in quartz sands were used as an ion-exchange resin for the adsorption of six PFAS (PFBS, PFOS, PFHpA, PFOA, PFHxS, and PFNA) in a low concentration range up to 100 μg/L and a high concentration range from 50 to 500 μg/L in a single and multi-solute system. The K_f value for PFOS in a multi-solute system is 404 L/g. At the start, the adsorption of PFHxS and PFNA increased at aqueous concentrations less than 5 mg/L and remained constant till 40 mg/L (3.89% PFNA and 23.6% PFHxS), then decreased at a higher equilibrium concentration (100 mg/L). Studies in both batch reactors and columns showed that at higher concentrations (60 mg/L total), longer-chain length PFAS, notably PFOS, were predominantly adsorbing. Adsorbed levels of all six PFAS were lower in the multi-solute system compared to those obtained from trials conducted with a single solute system [105].

To remove GenX, and two additional perfluorinated ether acids (PFEAs), researchers utilized a highly basic anion exchange resin (Purolite^® A860) in both surface and recycled water. Isothermal tests included dissolving 10–1000 mg of resins in 1 L of water sample with 0.1–100 g GenX and 5 mg C/L starting concentration of natural fluids and letting the mixture stand at room temperature for 24 h (equilibrium). Following 24 h of treatment, the resins were removed from the water with 0.45 m pre-rinsed syringe filters (Fisher Scientific, SLHV033RS, Nepean, ON, Canada) before the final PFAS concentrations were measured. Uptake is represented as a linear function of compound charge density, with the largest uptake occurring for the utmost charged portion (PFMOPrA > PFMOBA > GenX (p 0.05)). This is consistent with earlier research on IX, which found that organic molecules with larger charge densities were selectively eliminated. With a trend of 1.02–0.04 (R² > 0.98), they find that the intake of all PFEAS is linearly related to the analogous outflow of chloride ions. This data suggests that extremely basic macroporous IX resin (A860) absorbs GenX and other perfluorinated ether acids by a mechanism similar to that of ion exchange (IX), with just a little contribution from hydrophobic interactions and other physical variables [104]. Reversible addition-fragmentation chain transfer (RAFT) polymerization was used to create a wide variety of PFPE-containing polymers, both ionic and cationic, and then bonded onto magnetic iron oxide nanoparticles (IONPs, Fe₃O₄). The quaternized ammonium monomer of 2 and 9 and the degree of polymerization (DP) of oligo (ethylene glycol) methyl ether acrylate (OEGA), respectively, were found to provide the outstanding outcome in the separation of numerous PFAS, and P2-9+@IONPs, the grafted polymer sorbent consisting of the greatest percentage of cationic groups (39 wt%), was the most effective (Figure 4).

The addition of the quaternized ammonium groups dramatically increased the PFAS removals from all three quaternized polymer-grafted IONPs (P2-9+@IONPs, P4-6+@IONPs, and P6-3+@IONPs). It is believed that electrostatic attractions between the polymer’s quaternized ammonium group and the PFAS’s charged head group are responsible for this phenomenon, which is notably evident in the elimination of GenX and other short-chain PFAS. P2-9+@IONPs demonstrate quick capture of GenX with an outstanding removal rate, capturing >99% in 30 s and increasing to 99.5% in less than three minutes [108]. IX’s performance is highly impacted by the presence of co-contaminants in the feed solution, and the process generates secondary waste during the regeneration phase. Furthermore, the resin is sensitive to oxidation. Overall, ion exchange resin is a viable option for the efficient treatment of water contaminated with PFAS. However, choosing the right resin is essential, since long-chain PFAS have a higher affinity for the resin and need the use of specific resins for their removal. Furthermore, the resin has a limited capacity and must be renewed whenever it reaches saturation.

2.2. Membrane Separation Technology

Nanofiltration (NF) and reverse osmosis (RO) are two membrane-based treatment methods that depend on the elimination of PFAS by charge and size exclusion, or sorption onto the membrane polymers. PFAS rejection via membranes may be as high as 99% in certain cases, but their use is limited by energy requirements and a limitation of molecular selectivity. The summary of membrane filtration and reverse osmosis for PFAS removal is represented in

Table 3. Removal of PFAS via nanofiltration and reverse osmosis.

Membrane	Water Flux	Feed Solution	pH	Removal Efficiency	Duration (h)	Mechanism	References
BW30-RO		PFOA @ 10 µg/L	-	92%	6	-	[109]
RO		GenX @ < 0.01		>99%		-	[110]
NF		GenX @ 1 mg/L		99.50%	6	-	[111]
NF90		PFOS @ 100 ppb		96.40%	-	-	[112]
		PFOA @ 100 ppb		97.40%	-	-
NFG	1.4 × 10−5 m/s	PFOS @ 100 ppb	7.4	27.90%	24	-	[112]
		PFOA @ 100 ppb		55.60%	24	-
BW30		PFOS @ 5.0 × 102–1.6 × 106	4	>99.1	-	-	[113]
BW30		PFHxA @ 1.0 × 105	3.5–7.1	96.0 to >99.9	-	-	[114]
ESPA-2540		PFPeA, PFHxA, PFHpA, PFOA, PFPrS, PFBS, PFPeS, PFHxS, PFHpS, PFOS @ 0.3–58.4	7.0 ± 0.4	>99.0 (ALL)	-	-	[115]
ESPA-2540		PFPeA, PFHxA, PFHpA, PFOA, PFPrS, PFBS, PFPeS, PFHxS, PFHpS, PFOS)	7.0 ± 0.4	>99.0 (ALL)	-	-
XLE		PFHxA @ 1.0 × 105	3.5–7.1	>97.5	-	-	[114]
SW30XLE		PFHxA @ 1.0 × 105	3.5–7.1	>96.0	-	-
NF90		PFOA @ 10 µg/L		72%	5	-	[109]
		GenX @ 200 μg/L	-	97%	-	NaCl/methanol solution	[116]
Porous amine-functionalized membrane		PFOS @ ≤ 1.2 × 106 μg/m2		>99%	-	-
		PFOA @ ≤ 1.2 × 106 μg/m2		>99%	-	-
ASFPAN nanofibrous membrane		GenX @ 100 mg/L	4	35%	-	superhydrophilicity and the Coulomb force	[117]
ESPAN nanofibrous membrane				88%	-	hydrophobic interaction and dipole–dipole interaction
Five commercial membranes DK > NF90 > XN45 > NF270 > DL	≤0.1 L⋅m−2·h−1	PFCA @ 100 µg L−1	5.5 ± 0.06	67.6% to 95.8%	-	Hydrophobic interactions	[118]
				79.2% to >99.9%	-
		PFSA @ 100 µg L−1		79.2% to >99.9%	-
		FTS @ 100 µg L−1		69.0% to 97.2%	-
		PFEA @ 100 µg L−1		66.0% to >99.9%	-
PEMe		PFOS @ 1.0 × 103 μg⋅L−1	7	65.0–90.0	-		[119]
		PFOA @ 1.0 × 103 μg⋅L−1		35.0–90.0	-

.

The current treatments used combined approaches and hybrid membrane technologies for the elimination and destruction of PFAS via adsorption, ion exchange, and membrane separation. In membrane separation, nanofiltration and reverse osmosis are efficient techniques for PFAS removal from wastewater. Novel membranes consisting of nanomaterials were the most efficient in removing PFAS. The cost of membrane treatment and ion exchange has been found to be very effective. Irrespective of the chain length or molecular structure of PFAS, Ross et al. reported that RO and NF are very efficient at eliminating several of them, particularly PFAA intermediates [120]. They highlighted the significant challenges inherent in these methods, such as the high cost of large-scale implementation, the difficulty of dealing with suspended particles and water geochemistry, and the difficulty of keeping membranes healthy, especially in groundwater applications. Figure 5 displays the NF/RO membranes that have been experimentally indicated to be effective in the removal of PFAS from water.

Nanofiltration and reverse osmosis have been proven to be effective in the removal of per- and polyfluoroalkyl compounds from water, and their efficacy has been evaluated by [121]. The NF membrane used for the treatment of PFAS had the molecular weight cut-off (MWCO) and pore size in the range of 90–1000 Da and 0.3–2.1 nm, while the RO membrane was in the range of ≤100 Da and ~0.1 nm. The highest removal rates were shown by the most commercial NF membranes, about 90.0% to 99.0% for common PFOS and PFOA. On the other hand, removal by RO membrane showed 96% to 100% removal rates which is greater as compared to NF membrane. It was concluded that overall, the eradication of PFAS via RO and NF membrane depends upon size exclusion and electrostatic and hydrophobic interactions. The process of PFAS elimination is also influenced by the membrane characteristics and physiochemical properties of PFAS. Liu et al. reviewed the effectiveness of nanofiltration and reverse osmosis technologies for the removal of per- and polyfluoroalkyl substances from water. The NF membrane used for the treatment of PFAS had a molecular weight cut-off and pore size in the range of 90–1000 Da and 0.3–2.1 nm, while the RO membrane was in the range of ≤100 Da and ~0.1 nm. The highest removal rates were shown by the most commercial NF membranes about 90.0% to 99.0% for common PFOS and PFOA. On the other hand, removal by RO membrane showed 96% to ~100% removal rates which is higher than NF membrane. It was concluded that overall, the removal of PFAS by RO and NF membrane depends upon the size exclusion and electrostatic and hydrophobic interactions. The process of PFAS removal is also affected by the membrane characteristics and physiochemical properties of PFAS. In another study, Zhi et al. used five different types of commercial membranes to analyze their removal efficiency for three perfluoroalkyl ether acids (PFEA), seven PFAS, and two fluorotelomer sulfonates (FTS) [118]. Rejection rates for GenX (C5) were 11.6% greater compared to PFHxA (C5) on average (p < 0.05). When evaluating the effectiveness of the investigated membranes in removing PFOA and GenX, PFOA removal was shown to be more consistent. In one study, it was shown that powdered activated carbon (PAC) made from wood using 60 mg/L of dosage was able to successfully adsorb 80% of PFOA, >95% of PFOS, and 30% of GenX. Even at a dosage of 100 mg/L carbon, short-chain PFEA like perfluoro-3,5-dioxahexanoic acid (PFO₂HxA) were almost completely nonabsorbable. Short-chain chemicals (perfluoroalkyl-chain lengths of 3 and 4) including PFMOBA, PFBS, 4:2FTSA, PFMOPrA, and PFBA were effectively removed by DK and NF90 membranes. To treat certain types of PFAS, the NF method emerges as a more viable approach over GAC adsorption.

While the RO and NF are effective in removing PFAS, they do have certain drawbacks. Energy input is substantial, and the process may be costly to run. To minimize performance degradation due to fouling, the membranes need to be cleaned and maintained often. It is important to consider the high energy consumption and maintenance expenses connected with the technology. Instead of relying only on RO or NF technologies, it may be preferable to use these technologies as a pretreatment process in the treatment system for fluids with complex mixtures like groundwater and wastewater.

2.3. Fractionation Techniques

In the physiochemical process known as foam fractionation (FF), pollutants that are surface-active compounds may be extracted from a growing foam and purified (Figure 6). By compressing the accumulated foam, a waste stream rich in surface-active contaminants may be created. Foam fractionation seems ideally adapted for the isolation of these chemicals from surface and groundwater systems because of the considerable surface activity of many distinct kinds of PFAS molecules, with the compressed foam resulting in a low proportion of the remaining fluid waste stream. Recently, foam fractionation has received more attention, which has led to the discovery of novel uses for this technique [122]. Surface-active foam fractionation (SAFF) was used in a field study by Burns et al. (2021) to clean up PFAS in groundwater at an Australian location with a treatment capacity of up to 250 m³/day. The SAFF process was more effective at removing PFAS with relatively long fluorocarbon chains and greater air/water adsorption ratios because these PFAS had a stronger affinity for the bubble surface [123].

In another study, a bubble fractionation technique was used to analyze the influence of different environmental parameters including temperature, salinity, and aeration on PFAS removal in the absence of a co-surfactant. Before subjecting contaminated waters to high-energy PFAS destructive procedures, aeration removal may be a viable option as a non-destructive means of separation/concentration of amphiphilic (i.e., surface-active) chemicals. Aeration causes gas bubbles to develop in a liquid, creating a gas-liquid interface where the desired surface-active molecules may bind.

As the bubbles float to the top of the liquid column, they are easily isolated from the fluid being processed [124,125]. The FF technique has benefits over more traditional preconcentration methods like adsorption on AC or membrane filtration because of its minimal consumables use and resistance to complicated and changing water matrices [123,126]. Short-chain or non-amphiphilic PFAS are more difficult to remove since this method mainly effective for surface-active PFAS [127]. PFDS, PFHpS, PFHpA, and PFNS were likewise nearly completely eradicated after 60 min of aeration, whereas PFHxA, PFBS, and PFPeS were removed to varying degrees. At the conclusion of the 60-min aeration treatment, the total PFAS species were removed with efficiencies of 95.7% in the room-temperature condition, 97% in the low-temperature condition, and 93% in the high-temperature condition. Although aeration proved to be an effective method of PFAS treatment, the concentrated PFAS foamate that is generated must be further treated in order to degrade the PFAS compounds and remove them from waste streams [125]. Buckley et al. conducted a study wherein sodium dodecyl sulfate was utilized as a hydrocarbon surfactant to investigate the impact of salts on foam fractionation of PFAS species. It was discovered that the addition of salts improved the efficiency of PFAS removal due to an increase in surfactant adsorption at the air-water interface. However, it should be noted that the inclusion of extra chemicals may restrict the locations where the treated product water may be dumped [122]. Foam fractionation was utilized by Smith et al. to concentrate PFAS in groundwater, and then electrochemical oxidation (ECO) was employed to destroy the PFAS [127].

There was a 10 L min⁻¹ air flow rate, 20-min residence duration, and a 10% collected foam percentage. The system ran constantly for 9 h, and samples of influent, effluent, and foam (at 2, 4, 6, and 8 h) were taken for PFAS measurement. In a 20 L circulation tank with a combined total stainless-steel cathodic surface area of 9200 cm² and BDD anodic, 9 h batch tests were performed in triplicate for electrochemical oxidation. Ninety percent of PFAS were removed from groundwater and leachate during foam fractionation. The electrochemical degradation technique degraded PFAS by 84% after 540 min long-chain PFAS was eliminated more efficiently than short-chain PFAS (mean 22%) due to their greater air–water interface adsorption ratios [127].

Foam fractionation works well on its own, but when ozone is added, the process becomes much more efficient (Ozofractionation). To create ozofractionation, the ozofractionation catalyzed reagent addition (OCRA) procedure oxidizes organic pollutants chemically while simultaneously producing intense foam fractions that can be easily isolated from the final treated water product. Foam fractionation can efficiently remove PFAS in the liquid media because the perfluoroalkyl groups shift to the gas-fluid interface, and ozone oxidizes and degrades organic molecules. Although ozone (O₃) has only been explored in small-scale experiments for the cleanup of PFAS in water, it has been shown to achieve removal rates > 97% in a sensible procedure time [128]. Full-scale assessment employing ozofractionation for PFAS elimination proved to be successful, as reported by Horst et al. (2018). In order to meet stricter emission regulations, polishing treatment steps, including adsorption and filtration, were included [129]. O’Connor et al. utilized ozofractionation for the treatment of PFAS-contaminated marine fish hatcheries. OCRA is a revolutionary technique designed by the University of Newcastle and its industrial partner, Evocra Pty Ltd., for the removal of PFAS, acid mine drainage, and hydrocarbons from water. The OCRA technique is a successful fractionation method that controls the pH and oxidation-reduction potential of a solvent by using high quantities of ozone. Experiments conducted at both sites showed that commonly existing foam fractionation systems treating polluted saltwater with air or ozone could eliminate more than 90% of the PFOS and PFOA present. Ozone is widely employed in aquaculture systems today to disinfect saltwater, and research indicates that it is effective in removing PFAS. Nonetheless, ozone usage requires caution, and future studies are necessary to properly optimize its use in fish production [130].

2.4. Destruction Techniques

Destructive technologies are a kind of remediation that may, in most cases, decompose pollutants into safer byproducts. There are several destructive technologies that have been shown to be effective in degrading PFAS, with removal efficiencies ranging from 60% to 100%, including electric field-nanofiltration, electrochemical water treatment technology, direct current plasma, plasma-based technologies, catalytic ozonation, etc. A vast variety of PFAS compounds have been found to be destroyed and defluorinated to a great degree using hydrothermal alkaline treatment (HALT), mainly with NaOH. Compressed water with a high pH (>14), high pressure (25 MPa), and high temperature (350 °C) may be used to convert PFAS into innocuous fluoride salts if the kinetics are severe enough [131]. Although perfluorocarboxylic acids (PFCAs) are destroyed more quickly, perfluoroalkyl sulfonic acids (PFSAs) need lengthy residence periods (30 min) and large hydroxide loadings (e.g., 5 M-NaOH) in batch HALT reactors to attain greater percentages (>90%) of degradation and defluorination [132,133,134]. Pinkard et al. used three different concentrations of NaOH: 5 M-NaOH (200 g/L), 0.1 M-NaOH (4 g/L), and 1 M-NaOH (40 g/L) during the HALT. The inner temperature of the continuous reactor is maintained at 350 °C throughout operation. Changing the pump’s flow rate at a constant internal pressure of 25 Mpa causes a change in the reaction’s residence time from 1.6 to 10 min. This work seems to show that the residence periods needed to accomplish >99% annihilation of parent PFAS species may be reduced using the continuous flow HALT reactor setup. Residence periods of 10 min. resulted in >99% annihilation of the most resistant PFSA species, including PFHxS, PFPeS, and PFBS, whilst PFOS levels were reduced by >99.99% after only 1.6 min of processing. Based on these findings, it seems that continuous flow HALT technology may be expanded to practical throughputs [134].

Novel treatment techniques that can be effective against PFAS at room temperature and pressure are desperately needed. Yang et al. used a nonthermal ball mining technique for the destruction of PFAS. In a jar (100 mL) containing zirconium and stainless-steel balls, the PFAS-contaminated sediment samples collected from Schriever Air Force Base (Colorado Springs, CO, USA) were combined with co-milling reagents (KOH and boron nitride) and spun at a rate of 580 revolutions per minute for different time intervals. After that, the samples were analyzed using LC-MS/MS. After 6 h of treatment, almost 80% of the 21 PFAS were eliminated. It was found that the reaction pathways included both fluorination of boron nitride and piezo-electrochemical oxidation of PFAS [135].

Austin et al. contrasted the efficacy of continuous flow supercritical water oxidation (SCWO) for PFSAs and PFCAs [136]. The destruction seemed directly proportional to the rising temperature. At 610–650 °C, the destruction and removal efficiencies (DRE) were found to be more than 99.999%. Trace amounts of PFOS (1.66 g/L) were the only PFAS found in the effluent at T = 650 °C, which is equivalent to a DRE of >99.999%. Keep in mind that at lower working temperatures, PFHpS seems to have a lesser DRE compared to PFOS. It was discovered that both PFSAs seem to be more persistent than PFCAs. This is consistent with the comparison of PFOS and PFOA performed by [137]. Li et al. have described the decomposition and mineralization of PFOS and PFOA across a wide temperature spectrum (T = 410–650 °C). The PFOS DRE was found to be more than 99.999% at a residence duration <30 s and T = 650 °C. In the wastewater discharges from gaseous channels, PFOS are converted into a pool of PFCAs, including 1H perfluoroalkane, at lower temperatures (T = 420–600 °C). It is important to deal with the toxic byproducts formed during and after the treatment. Negative public health and environmental repercussions occur if toxic liquid and gaseous wastes are not contained.

The possible drawbacks of traditional membrane filtration may be overcome by incorporating electrical activity into the separation process. As PFOA has a strongly electronegative carbon-fluorine bond, its calculated pKa shows that it is negatively charged in water across common pH ranges. This makes it fair to hypothesize that PFOA would be impacted by an electric field. The literature reports that putting a direct-current electrical field over a microfiltration membrane dramatically increases PFOA and PFOS elimination from 0% to 70% [138,139]. The elimination of PFOA was investigated using a crossflow membrane cell constructed from a porous SnO₂-Sb anode and a porous titanium cathode in an electric field nanofiltration [140]. A cross-sectional element mapping overlay of Sn and Ti, proving that the SnO₂ layer was effectively deposited on the Ti substrate and demonstrating its coherence and density.

Researchers used a voltage gradient between 0 and 90 V while maintaining a constant transmembrane pressure of 4.14 bar (60 psi). From 0 to 30 V, the electrical voltage increased the PFOA rejection from 45% to 97%. Compared to non-saline feed, the saline feed has a reduced PFOA removal rate with inconsistent filtration effectiveness, ranging from 55% to 80%, which might be due to the bubble formation during water electrolysis, deteriorating the electric field force. Expanding the applicability of electrochemical water treatment technology, electric field-assisted nanofiltration offers a revolutionary separation strategy for efficiently eliminating ionized organic contamination [140]. Microwave power supply, direct and alternate current, radiofrequency sources, and the pulsed discharge technique are only a few of the ways that non-thermal plasma may be generated for use in water and wastewater purification. The removal of PFOA and PFOS by direct current plasma has been stated to be 100%. Hence, plasma-based technologies are highly efficient in filtering PFAS out of both natural and artificial water supplies. Its efficiency, though, suffers when it encounters inorganic and organic contaminants from outside sources. In addition, the plasma technique produces harmful by-products. Notwithstanding its limitations, this technology requires more investigation on a large-scale, real-world scale [141].

Lashuk et al. evaluated the catalytic ozonation removal potential for five different PFAS including GenX in the presence of four catalysts fabricated using titanium in different ratios of butaoxide, mixed with t-butanol (WO₃/TiO₂ catalyst) [142]. Photocatalytic ozonation removed 4–24% of PFAS, whereas photolysis destroyed 4–13%, ozone photolysis eliminated 7–16%, and photolysis removed 2–4%. In general, after 4 h, 6:2 Fluorotelomer sulfonic acid (6:2 FTSA) and PFOA were removed at higher rates than PFHxS, PFBS, and GenX. When comparing the short-chain PFAS, photocatalytic treatments removed approximately twice as much PFHxS (6-carbon chain) as PFBS (4-carbon chain). A developing issue due to their increased usage in industry and presence in the environment, poorer removal percentages for shorter-chain PFAS have been reported by researchers. GenX, the new PFOA replacement material, proved very resistant to all treatments. This is in line with the latest results of [42], who found that the CF3 group on the alpha carbon of GenX acts as a protective group, making it resistant to advanced oxidative processes [142].

The most extensively researched approaches for PFAS destruction are electrochemical advanced oxidation processes (EAOPs), which produce highly susceptible hydroxyl radicals (OH•) with a high redox potential (E = 2.8 V versus SHE). They may generate immiscible alcohol compounds to break C–F bonds. Due to its efficacy and simplicity in current water treatment facilities, anodic oxidation is a popular EAOP [143]. Using ultraviolet light and persulfate, [144] investigated the oxidation of hexafluoropropylene oxide tetramer acid (HFPO-TeA), reporting that HFPO-TeA depolymerizes to Hexafluoropropylene oxide trimer acid (HFPO-TA), which in turn depolymerizes to HFPO-DA.

The effectiveness of boron-doped diamond (BDD) anodes in the breakdown and defluorination of HFPO-DA was studied by [20]. Investigations involving ECO were conducted in a polypropylene electrochemical cell with a single chamber. The operating anode was a boron-doped diamond electrode with an efficient surface area of 22.6 cm², and the cathode was a platinum mesh set up adjacent to the anode on each side, with a distance of 10 mm between the electrodes. In order to determine the fluoride content, an ion-chromatograph system (DX-120, DIONEX, Lane Cove West, NSW, Australia) was used. This system included a separation column (IonPak AS12A, 200 mm, and 4 mm) and a suppressed conductivity detector. The flow rate was adjusted to 1.3 mL/min and the mobile phase was an aqueous solution of 1 mM Na₂CO₃ and 1 mM NaHCO₃. At 4 h, the removal rate was nearly comparable for the two electrolytes, at 91% (Na₂SO₄) and 93% (NaClO₃). The observed reaction rates (kapp) in the Na₂SO₄ and NaClO₄ electrolytes were 0.0106 and 0.0113 mL/min, respectively, indicating that the process followed pseudo-first-order kinetics.

After 4 h, defluorination efficiency was 80% in Na₂SO₄ electrolytes and 84% in NaClO₄ electrolytes. Breakdown of HFPO-DA after 4 h was 96%, 92%, and 81% at 0.005, 0.01, and 0.05 M electrolyte concentrations, respectively (Figure 7,

Table 4. Efficiencies of different destructive technologies for PFAS removal.

Removal Method	PFAS Type	Initial Concentration	Additives	Removal Efficiency (%)	Reaction Duration	Defluorination Efficiency	Mechanism	References
Nanoscale Zero-Valent Iron (nZVI) on 2D reduced graphene oxide (rGO) nanosheets	PFOA	200 μg/L	-	39	10 min	-	-	[145]
	PFOS		-	85
HALT	GenX	20 μg/L	-	100	2 h	-	-	[146]
	F–53B	20 μg/L	-	12.9 ± 3.4	2 h	-
	ADONA	20 μg/L	-	100	2 h	-
HALT	AFFF	3200 mg/L	NaOH	99.95	2 h	-	-	[134]
HALT	AFFF	-	NaOH	−100	1.5 h	98	-	[132]
HALT	PFOS	50 mg/L	NaOH	−100	1.5 h	>80	-	[147]
HALT	AFFF (PFAS mixture)	38,714.5 μg/L	air	99.37	10 s	-	-	[148]
Boron-doped diamond anodes	GenX	15 mg/L	-	57%	1 h	>80%	Oxidative degradation, direct electron transfer	[20]
	PFOA			38%
Electric field-assisted nanofiltration	PFOA	0.5 mg/L	Na2SO4	97.5%	48 h	90%	1-electromigration of PFOA away from the membrane, 44 2-electro-absorption of PFOA on the electrodes, 45 and 3-membrane fouling enhancement	[140]
Photocatalytic degradation	GenX	100 µg/L	-	99%	1 h	-	-	[149]
piezoelectric-material-assisted ball milling	PFOS	-	-	>80%	2 h	100%	piezoelectric oxidation pathway	[135]
				>100
	PFOA

). The rate of HFPO-DA breakdown was shown to be determined by direct electron transfer. It is suggested that HFPO-DA degrades via the creation of three intermediates before finally mineralizing to CO₂ and F (Figure 7) [20]. Unfortunately, the high upfront expenditures and ongoing operational and maintenance expenses of these technologies have prevented their widespread implementation on a practical scale.

3. Removal Mechanism

Understanding the mechanism of PFAS adsorption and destruction helps with both the assessment of removal techniques for prospective applications and the development of novel, more efficient removal techniques. According to Smaili and Ng, 34 investigations found PFAS sorption to be dominated by hydrophobic contacts, 19 by electrostatic interactions, and 48 by a mix of both [150] (Figure 8).

Other interactions, including hydrogen bonding, fluorine-fluorine interactions, and ligand exchange (although weakly), were shown to influence the adsorption of PFAS in eight more investigations. Research was conducted by Ma et al. to assess the influence of iron minerals in laboratory-scale parallel-constructed wetlands (CWs) for the elimination of PFOA and PFOS [151]. To remove PFOS and PFOA from CWs, iron minerals of different shapes and sizes were used as substrates, with the goal of optimizing microbial degradation and substrate efficiency. The maximum removal of PFOA and PFOS (57.2% and 63.9%) was exhibited by MC2 (Gravel:Magnetite, 1:2). The concentration of PFOA and PFOS in the effluents were decreased when iron minerals were added to CWs, relative to the control group. Their elimination is aided by electron transfer and denitrification, both of which are favorably linked with iron content (Figure 9) [151,152]. Moreover, because of the increased translocation factor of PFOA in CW, its removal was more successful than that of PFOS.

Two main pathways exist for emerging contaminants biodegradation by microalgae: metabolic degradation, in which the emerging contaminants act as a carbon source or electron giver/acceptor, and co-metabolism, in which the emerging contaminants are destroyed by enzymes activating other substrates [153].

Adsorption of PFAS onto carbon nanotubes (CNTs) is a complex process that has not yet been fully elucidated. Du et al. concluded in their review that hydrogen bonding interaction is not likely to be involved in the adsorption process of PFCAs on CNTs [70]. Recent research reveals that charge-assisted hydrogen bonding may react with the carboxyl group of the PFCAs and the functional group on the CNT surface to adsorb ionizable molecules on CNTs.

RCOO⁻ +H ++ -O-CNTs (RCOO…H…O-CNTs)

The carboxyl, hydroxyl, and amine functional groups on the CNTs surface also attract water molecules which form water molecule clusters on the CNTs surface, which could be the reason for negligible involvement of hydrogen bonding in PFCAs adsorption on CNTs due to the competitive adsorption. The hydrophobicity of the PFAS alkyl chain is proportional to the amount of PFAS adsorption on CNTs. PFOA, PFBA, and PFHxA all contain CF chains with different numbers of carbons (eight, four, and six, respectively). Adsorption studies using CNTs reveal that after 15 h, over 95% of PFOA had been absorbed using single-wall carbon nanotubes (SWCNTs), but PFBA (7.5%) was hardly eliminated [154,155,156]. Adsorption on carbon nanotubes increases with hydrophobicity and PFAS chain length [70]. In water, PFAS mostly exists in anionic forms and CNTs have anionic functional surface groups causing electrostatic repulsion and resisting PFAS adsorption. However, the attractive hydrophobicity of the CNTs surface is thought to outweigh the repelling electrostatic interactions [157]. The lack of surface hydrogen in most CNTs prevents them from forming H-bonds with the fluorine atoms in PFAS [156].

KD for perfluorocarbons of the PFCA and PFSA series increases with chain length because of the larger dispersion forces and solvophobicity that the perfluorinated chain may exert at the interface [158]. According to past research of PFAS adsorption on AC, the sulfonates adsorption is stronger than the carboxylates [63]. The different group characteristics of the charged CO₂ and SO₃ units explain why sulfonate is more attractive than carboxylate [110]. As compared to the carboxylate ion, the sulfonate ion has lower solvation because its charge is more delocalized. The sulfonate is preferentially favored in its surface-dispersion interactions, as dispersion is related to polarizability and the total of the polarizabilities of the atoms in their ground state [63]. In a study, Min et al. fabricated functionalized periodic mesoporous nanosilica (PMO) with different amine contents using a condensation method for the removal of a short-chain PFAS, perflourobutanoic acid (PFBA). The functionalized PMO material with the maximum fluoroalkyl group loading of 1.05 mmol/g (PMO-NHCF-3, removed PFBA with little interference from coexisting anions. Adding a fluoroalkyl group to an aminated organic silica may enhance nonelectrostatic interactions with PFBA, which can then increase PFBA adsorption onto bifunctional PMOs while decreasing inorganic anion (IA) competition. The wide contact angle of n-hexadecane on PMO-NHCF-3 (128°) further demonstrated the material’s significant oleophobicity, which is a unique feature of the fluoroalkyl group. As PFBA has a distinctive fluorous chemistry and a stiff structural C-F skeleton, fluorophilic interactions could further improve the attraction among PFBA and the bifunctional PMOs [159,160,161].

Mantripragada et al. produced and compared the removal efficiency of two types of nanofibrous membranes polyacrylonitrile (ESPAN) and amidoxime surface functionalized ESPAN (ASFPAN) for the removal of GenX from water [117]. The water contact angle of the ESPAN is 11 ± 52, making it hydrophobic. GenX adsorption on ESPAN showed hydrophobic contact regardless of the negative surface charges. Because of the high electronegativity of F, the CF chain of GenX may include strong dipoles of ^(δ+)C−F^(δ−), whereas polyacrylonitrile (PAN) molecules may have strong dipoles of ^(δ+)C−N^(δ−). These strong dipole-dipole forces might be the cause of GenX adsorption on ESPAN. More concentrations of amidoxime functional groups are present in ASFPAN molecules than in PAN, indicating that they may be positively charged in acidic environments. Adsorption of GenX onto ASFPAN molecules may be facilitated by electrostatic contact (Coulomb forces) between positively charged sites >C−N⁺< on ASFPAN molecules and the negatively charged COO⁻ groups on GenX. The surface hydrophilicity and electrostatic interaction contribute to a nearly twofold improvement in GenX eradication rate by the ASFPAN versus ESPAN [117].

4. Cost Analysis

It is estimated that for every 1 g of PFAS removed from a polluted site, the Global warming potential (GWP) increases by 83–122 kg of carbon dioxide equivalents (eq), and the cost increases by >75 United States dollars each gram of PFAS treated [75,162]. Murray et al. did the cost analysis of two PFAS treatment techniques, AC and IX. Treatment costs were compared to 10% and 50% breakthrough based on media choice and treatment goals by conducting a sensitivity analysis of operation and maintenance (O&M) expenses for GAC and IX projects [163]. Calculations show that compared to average-performing GAC, which needs a changeout around every other month, PFAA treatment by IX is at least 20% less costly (given by levelized media cost) throughout all unit cost points studied due to 16 times longer changeout period.

To reach a point of equity between GAC and IX, effectiveness must improve by 25% in GAC to reduce changeouts or unit cost must fall to less than $2.75/kg, while performance must decline in IX, leading to an increase in unit cost of more than $20.35/kg. There is still needed to develop efficient cost-effective disposal techniques for the used adsorbents for PFAS removal, for now, the AC and IX are piled up in the landfills. However, the value of porous carbon-based materials is far cheaper than that of many other absorbents. Chemically activated carbon made from nutshells costs roughly USD 1.82 per kilogram to produce, according to recent research [164]. Another study reported the chemical and physical activation of AC was between $0.96–4.22 and $0.96–1.54 [165,166]. A study has been conducted on the cost analysis of AER and GAC for GenX and PFAS. To treat water polluted with GenX or PFOA using GAC, the yearly cost per home is $224 or $85. There is no difference between GenX and PFOA in terms of direct or total capital expenses related to AE resin (Type I strong base polystyrenic gel). Operating and maintenance expenses for treating GenX and PFOA with AE resin are similar (US$156,594 and US$157,482, respectively) per year. Costs of operation and maintenance fluctuate somewhat from year to year owing to differences in GenX and PFOA removal rate of AE resin [22].

The GAC filters were installed in a Ridgewood, New Jersey car treatment plant which costs $3.5 million for installation and treats 1 million gallons/Day [167]. In comparison to synthetic materials (US$1500 ton⁻¹ for activated carbon), utilizing biomaterials (US$246 ton⁻¹ for activated biochar) could aid in cost savings [168]. The Sweeney water treatment plant was upgraded by adding 8 GAC filters in Wilmington, North Carolina having an installation cost of $46 million and an operational cost of $2.9 million for the treatment of 44 million gallons/Day [169]. As compared to other water treatment methods, RO is the most affordable ($2 m⁻³), even when the average cost of power is factored in (8.14 cents KW h⁻¹) [170]. A new RO plant was installed in West Morgan-East Lawrence Water Authority, Decatur, Alabama for $30 million and costs $1 to clean 10 million gallons/Day [167,171].

Kanchanapiya and Tantisattayaku worked to analyze the excessive cost of removal of PFAS by using reverse osmosis from landfill leachate in Thailand [172]. Pretreatment and evaporation ponds were part of the two-pass reverse osmosis (RO) system used in this research. Using the water balance approach, they calculated the total volume of 111 sanitary landfills and analyzed the project’s financial viability using the net present cost (NPC) method. The findings indicated that the annual LL volume was about equal to 17.5 cubic meters. The RO system’s NPCs were calculated to be $577.9 million USD for a setup that included an evaporation pond, and $391.9 million USD for one that did without. The systems with an evaporation pond had a treatment unit cost ranging between $1.72 and $2.71 per cubic meter, whereas those without had a cost of between $1.06 and $2.09 per cubic meter.

The EPA presents the treatment technologies (GAC, IX, and RO) and costs for PFAS removal from water in an Annual Conference 2020. The total annual cost for PFAS treatment using GAC, RO, and IX is approximately $48,000,000, $49,000,000, and $50,000,000, respectively for more than 2000 households. Estimated laboratory testing expenses for PFAS are $250 per sample. A tank having 30 lbs of GAC (1 cu ft) for whole house water costs $539 whereas two large tanks for whole house water filter having 240 lbs of GAC costs $3990. A small GAC system weighs 200 lbs and costs $1200 before installation, and RO modification for Point-of-Entry use weighs 150 lbs and costs $2000 before installation [173].

Cost-benefit analyses of oxidation-reduction, sonochemistry, and biological technologies are extremely scarce, and further study is needed to make these approaches scalable. The processes of conversion and the viability of these approaches for actual wastewater need to be demonstrated experimentally in more depth. It is well-documented that the use of multimodal approaches, such as the use of dual-frequency systems for sonochemistry, may reduce expenses by 23%, from $10 m⁻³ to $7.5 m⁻³ [166,170].

5. Conclusions and Future Recommendations

An in-depth analysis of the current developments in PFAS removal and degradation technologies is reported in this review article. Since the ban on longer-chain PFAS chemicals, investigations show that the shorter-chain varieties have become more prevalent in the environment. The increased amount of PFAS have been detected in the effluents of WWTPs compared to influents which is possibly due to the breakdown of PFOA and other long-chain precursors into shorter-chain derivatives. Our analysis of available treatment options suggests that short-chain PFAS are more difficult to eliminate than long-chain PFAS. This disparity may be attributable to the fact that ionic interactions are more important for the sorption of short-chain PFAS compared to hydrophobic ones for long-chain PFAS. The MOFs exhibited higher adsorption capacity compared to biochar and AC. Reverse osmosis and nanofiltration are both found to be effective techniques for PFAS removal, however, the need to clean or change the filter and degradation of the membrane is costly. The destruction technique is the most effective one, but its use is limited due to the formation of PFAS precursors during treatment techniques. Therefore, it is recommended to use combined treatment techniques for PFAS removal. Current studies have yielded several suggestions on how to proceed with PFAS removal in the future.

No research has been conducted on the use of MOF for the elimination of GenX. Larger surface areas for adsorbents may be achieved in metal-organic frameworks due to the presence of positively charged metals in their crystallized framework and the uniformity of their pore structure. Due to their electrostatic and hydrophobic interactions, they may be excellent fits for eliminating GenX.

Short-chain PFAS were less likely to be adsorbed by activated carbon than long-chain PFAS. The AC must be tailored via various methods and examined to remove short-chain PFAS and GenX.

New PFAS compounds, including fluorotelomers, require further study in wastewater because of their widespread use in industries like fluorinated impregnations, metal plating, oil production, and food packaging.

Future research might focus on optimizing electrode design and flow-cell arrangement to boost electrosorption kinetics and increase mass transfer in electrochemical reactors. Further mechanistic research is required to fully understand the mechanisms involved in short-chain degradation using a variety of redox-electrode materials.

Byproducts from several PFAS cleanup methods may be hazardous to human and environmental health. There should be more investigation into the hazards of various PFAS-removal technology.

It is challenging to create analytical methods that might identify the kinds and amounts of PFAS since the molecular identities of several PFAS on the worldwide market are still uncertain and PFAS are found in the environment at extremely low quantities.

Future research should include evaluating a combination of PFAS instead of single compounds due to the heterogeneity in degradation efficiency amongst the different PFAS studied.

The destruction techniques with combined treatment technique targeting the elimination of long and short-chain PFAS is suggested after reviewing the relevant technologies. The implementation expenses of such a strategy, though, continue to be a barrier to its widespread use. Consequently, it is essential to reduce the overall cost of the adsorption process, which includes the cost of regeneration.