Poly- and per-fluoroalkyl substances in water: Occurrence, analytical methodologies, and remediations strategies: A comprehensive review
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Nompumelelo Malatji
Abstract
Poly- and perfluoroalkyl substances (PFASs) are fluoro-organic compounds comprising thousands of anthropogenically produced chemicals with various industrial and consumer applications. This review compiles recent information on the sources, occurrence, and health effects of PFAS in aquatic environments. Secondly, as a primary requirement for assessing the PFAS concentration in water, this review systematically summarised the analytical methodologies (sample preparation and analytical detection techniques) for PFAS. Furthermore, health risks associated with PFAS in water are outlined. Finally, researchers worldwide have investigated the strategies for the remediation and elimination of PFAS from water. Previous studies have shown that PFASs are present in various water bodies with the highest concentration detected in Germany (94–4,385 ng·L−1 in river and drinking waters). The findings of this review further revealed that solid-phase extraction techniques were the most preferred for sample preparation compared to liquid–liquid extraction techniques. Solid-phase extraction technique improved the limit of detection and the limit of quantification of many analytical techniques to 0.010–1.15 and 0.030–4.00 ng·L−1, respectively. For PFAS remediation, the adsorption method and chemical oxidation using heat-activated persulfate and photochemical oxidation were the most used techniques. The most studied water matrices were drinking, river, groundwater, wastewater, and modelled ultra-pure water. The most used detection technique was found to be liquid chromatograph-tandem mass spectrometer (LC-MS/MS).
1 Introduction
Poly- and perfluoroalkyl substances (PFASs) are synthetic organic compounds whose hydrogens are either partially or fully replaced by fluorine atoms [1]. They can be categorized into non-polymer and polymer, as illustrated in Figure 1. The non-polymer PFAS include perfluoroalkyl acids (PFAAs), perfluoroalkane sulfonyl fluoride, perfluoroalkyl iodides, and per and polyfluoroalkyl ether-based substances [2]. Polymeric PFAS includes perfluoropolyether, fluoropolymers, and side-chain fluorinated polymers [2–5].
Classification of PFAS.
PFAS consists of a carbon chain of varying carbon lengths. They are categorized as either polyfluoroalkyl or PFAAs, wherein hydrogen atoms are either partially or completely substituted by fluorine atoms, as shown in Figure 2 [6]. Charged functional groups such as sulfonic or carboxylic acid are also attached at one end of the carbon chain and are responsible for the chemical stability of PFAS [7]. The resulting PFAS are classified as perfluoroalkyl sulfonic acids (PFSAs) or perfluoroalkyl carboxylic acids (PFCAs) [7–9]. Other forms of PFASs such as the ether and phosphoric acid-based compounds are rarely discussed or not studied in the literature. PFCAs and PFSAs have a low vapour pressure, and their lipophilicity increases with the increasing C–F chain length [10,11]. Since the C–F bond is the strongest covalent bond, PFAS is reported to be thermally stable [12]. In addition, PFCAs display anionic properties at environmentally relevant pH, based on their acid dissociation constant (pKa) [13]. PFAAs are reportedly derived from polyfluoroalkyl acids; the non-fluorinated bonds allow a pathway for degradation [13].
Perfluoro sulfonic acid with complete (a) and partial (b) fluorination.
PFAS displays both hydrophobic and lipophobic properties responsible for the low polarizability of fluorine atoms [11,14]. Also, as the hydrogen substitution by fluorine and the carbon-chain length increases, PFAS becomes more chemically inactive [15]. In contrast, their solubility in water decreases with the increasing carbon-chain length [16]. According to Jansen and Warming [17], sulfonic acid-based PFAS are more likely to bioaccumulate than carboxylic acid-based PFAS with the same chain length due to their higher acidity [17]. This could be attributed to sulfonic acid having two double-bonded oxygens that allow it to participate in resonance, which spreads its negative charge over a larger space, thus increasing acidity. However, carboxylic acid has only one double-bonded oxygen, making sulfonic acids much stronger than their carboxylic equivalents [17,18].
The most widely used and detected PFAS in water are non-polymeric long-chain perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) [19]. These PFAS are toxic to the fauna and flora and are tenacious in the environment. In addition, PFAS tends to bioaccumulate in environmental compartments and living organisms [20,21]. As such, their effect on human health [21], animals [22], and plants [23] continues to be investigated.
Therefore, the main objectives of this review are to 1) summarize the sources and occurrence in aquatic environments; 2) discuss the analytical methods used for the identification and quantification of PFAS in water; 3) review the ecotoxicological risk assessment (ERA) of PFAS towards humans, plants, and animals; and 4) highlight the various technologies used for elimination and treatment of contaminants.
1.1 Application, sources, and transportation pathway of PFAS
An overview of the application, sources, and transportation of PFAS is illustrated in Figure 3 [24]. Due to their outstanding water, oil, and dust repellence properties, PFAS have been used in various fields, such as the textile, manufacturing, mining, and packaging industries, as shown in Figure 3 [25,26]. Incorporating PFAS in fabrics, furniture, and many other goods as repellents improves material durability and increases the lifetime of products [20,25,26]. In the mining industry, fluorinated surfactants are used as hydrocarbon foaming agents in drilling fluids, which helps to reduce the amount of fluid lost during drilling and the potential formation of damage [27]. They also increase the wetting of sulphuric acid or cyanide used to leach ore in gold and copper mines to enhance metal recovery [27]. The packaging industry uses PFAS as a proofing agent when manufacturing take-out paperboard containers, fast-food wrappers, pet food bags, microwave popcorn bags, etc., to prevent the leaking of oil and grease [28–30]. PFAS are surfactants in the oil industry to enhance oil recovery and production [27,31]. They also manufacture aqueous film-forming foam (AFFF) used by fire and military departments to suppress fire [32]. Another fast-growing application of PFAS is their use in manufacturing non-stick cookware and bakeware [27,33].
![Figure 3
Application of PFAS, their sources, and how they are transported into the environment and humans. Reproduced with permission from the study by Araújo24et].](/document/doi/10.1515/revac-2023-0064/asset/graphic/j_revac-2023-0064_fig_003.jpg)
Application of PFAS, their sources, and how they are transported into the environment and humans. Reproduced with permission from the study by Araújo24et].
Although PFASs have proven beneficial in many applications, their biggest problem is that when humans ingest them, they accumulate in the body and cannot be metabolized. Literature suggests that consumption of PFAS-contaminated foods, drinking water, and dust from PFAS-treated houseware are the common pathways for human exposure [34,35]. In drinking water, PFAS may result from inadequate wastewater treatment due to the lack of proper design in wastewater treatment plant (WWTP) techniques [35]. In underground and surface water systems, short-chain PFAS are easily transported and introduced from landfills, firefighting activities, mining applications, manufacturing, etc., by runoffs and leaching due to their small size [36].
In addition, Szabo et al. suggest using treated wastewater for irrigation, thus preventing the introduction of PFAS in groundwater [37]. Another study indicated that the degradation of consumer products containing PFAS presents their precursors into the environment [38]. For instance, PFAS-treated waste products disposed of at landfills undergo mechanical breakdown with time. The broken-down PFAS particles are then carried with the leachate into the underground waterbed [39]. Finally, municipal or private groundwater sources could be polluted by directly discharging industrial effluents [40]. For example, small to medium industries such as car wash, food processing, battery, tanking, and chemical manufacturing around the Leeuwkuil WWTP have increased their water capacity by up to 116% from their effluents [40].
2 Global concentrations of PFAS concentration levels in water systems
The increasing detection of PFAS in various water systems worldwide proves their predominant application despite their ban, especially in countries such as the United States and Europe with standard regulatory measures [41]. At the same time, the problem of PFAS appears to be encountered mainly by developed countries. It is a “silent problem” in developing and undeveloped countries, possibly due to the lack of facilities, good techniques, and funds to investigate and quantify their presence. Therefore, exploring these chemicals is important, considering the increasing global export/import activities (goods, clothing, and food) to meet human demand. The results of some PFAS studies conducted worldwide are depicted in Table 1. Multiple PFAS have been identified in various water systems globally in the current decade. For example, a study by Groffen and coworkers investigated 15 types of PFAS in the Vaal River (South Africa), and their average limit of quantification (LOQ) concentration was 38.5 ng·L−1 [41]. This concentration complies with the United States Environmental Protection Agency (USEPA) regulatory limit of 70 ng·L−1 for individual/combined concentrations of PFOA and PFOS [56]. In addition, among the detected PFAS, PFOA and PFOS were dominant [42]. A similar study was conducted at Roodeplaat and Hartbeespoort, where perfluoroheptanoic acid (PFHpA), perfluorohexane sulfonic acid (PFHxS), PFOA, and PFOS were the most dominant PFAS in these water systems. The obtained concentrations were between 1.38 and 346.32 ng·L−1 at Hartbeespoort and between 2.31 and 262.29 ng·L−1 at Roodeplaat Dam [42]. The average concentrations of all PFAS were higher in the Roodeplaat Dam (0.14–89.04 ng·g−1) than in the Hartbeespoort Dam (0.03–31.37 ng·g−1). The higher readings in the Roodeplaat Dam can be attributed to the influx of water from the Pienaars River, which contains runoff from the Baviaanspoort sewage works and runoffs from Mamelodi township [42]. The prevalence of more PFSAs than PFCAs around both dams suggested that most activities conducted in these areas use more PFSA-containing products. Another study was conducted at the Plankenburg River (South Africa), where the lowest PFOS (<0.06–12.4 ng·L−1) and PFOA (12.8–62.6 ng·L−1) concentrations were detected. In that study, PFOA and PFOS concentrations were higher at the human settlement and industrial sampling locations, indicating that sampling points significantly (p < 0.05) influenced the PFAS concentrations detected. Contamination in these areas was reported to have resulted from human activities such as poor sewage removal, agricultural activities, improper management of solid waste, poor drainage systems, untreated industrial effluents, and poor housekeeping of recreational facilities, making the two locations point as sources of PFAS near the Plankenburg river [46]. The analysis of all the studies reported in South Africa, according to Table 1, showed that PFOA exists in South African water systems at higher concentrations than PFOS.
Global concentrations of PFAS in various environmental matrices
| Country | Water sources | Conc. level (ng·L−1) | Type of PFAS | Extraction method | Adsorbent type | Detection method | Ref |
|---|---|---|---|---|---|---|---|
| South Africa | Vaal River | 38.5 | PFOA and PFOS | SPE | WAX cartridge | HPLC-MS/MS | [41] |
| South Africa | Hartbeespoort Dam and Roodeplaat Dam | 1.38–346.32 | PFOA, PFOS, PFHxS, and PFHpA | SPE | — | HPLC-MS/MS | [42] |
| Ghana | Kakum River, Pra River, and drinking water | 197–398 | PFOA, PFOS, PFHxA, PFDA, and PFPeA | SPE | WAX cartridge | HPLC-MS/MS | [43] |
| USA | Treated drinking water and source water | 19.5 and 21.4 | 17 PFAS | SPE | WAX cartridge | HPLC-MS/MS | [44] |
| South Africa | Eerste River, Salt River, and Diep River | 146–390 and 23–182 | PFOA and PFOS | SPE | — | HPLC-MS/MS | [45] |
| South Africa | Plankenburg River | 12.8 ± 4.24 and 62.62 ± 4.86 | PFOA and PFOS | SPE | HLB cartridge | UPLC-QTOF-MS | [46] |
| France | Raw and treated tap water | >100 | PFBS, PFHxS, PFHpA, PFOA, PFBA, PFPeA, PFOS, PFHxA, and PFNA | SPE | WAX cartridge | HPLC-MS/MS | [47] |
| France | River Seine | 2–105 | PFPeA, PFHxA, PFHpA, PFOA, PFNA, PFDA, PFUnDA, PFDoDA, PFTrDA, PFTeDA, PFHxS, PFHpS, linear PFOS, PFDS, FOSA, MeFOSA, EtFOSA, and 6:2 FTSA | SPE | Graphite cartridge | HPLC-ESI-MS/MS | [48] |
| Germany | Rhine, Ruhr, and Moehne Rivers and drinking water | 94–4,385 | PFOA, PFOS, PFBA, PFHxA, PFHpA, PFDA, PFNA, PFUnDA, PFDoDA, PFBS, PFHxS, and PFPeA | SPE | — | HPLC-MS/MS | [49] |
| Germany and Spain | Treated effluent wastewater, river water, tap water, and bottled mineral water | 0.04–258 | PFBA, PFHA, PFOA, and PFOS | SPE | — | HPLC-MS/MS | [50] |
| Sweden | Uppsala drinking water | 25–45 | PFOS and PFHxS | D-SPE | Graphitized carbon cartridge | UPLC-ESI-MS/MS | [51] |
| Spain | Drinking water | 1.81 and 2.40 | PFOS and PFOA | SPE | WAX cartridge | UPLC-ESI-MS/MS | [52] |
| Spain | Treated drinking water | 3.0–21 and 4.2–5.5 | PFOS and PFOA | LLE | — | UPLC-MS/MS | [53] |
| Faroe Islands | Raw and drinking water | 0.75 | PFBA | SPE | WAX cartridge | UPLC-MS/MS | [54] |
| China | Hun River | 2.12–11.3, 2.68–9.13, and 0.40–3.32 | PFHxA, PFOA, and PFOS | SPE | HLB and WAX cartridges | UPLC-ESI-MS/MS | [55] |
In Ghana, the first report on PFAS, their detection, and concentrations was for tap drinking water and river water, and it was conducted by Essumang et al. The group detected five types of PFASs in the Kakum and Pra Rivers with mean concentrations of 280.80 and 397.63 ng·L−1, respectively [43]. The mean concentrations of PFOS and PFOA in tap water obtained from the Pra River and Kakum River were 97.5, 102.6 ng·L−1 and 89.7, 106.8 ng·L−1 [43]. These concentrations clearly show the limitations of traditional water treatment methods for eliminating PFAS in drinking water. According to the study, contamination in the Pra River resulted from agricultural activities and the mine waste from illegal small-scale gold mining activities [43]. Another study in the United States analysed and detected PFAS in 50 water samples, 25 of them from drinking water treatment plants (DWTPs) near two rivers [44]. Only one sample from the DWTP exceeded the USEPA guideline value of 70 ng·L−1. The most dominant PFASs in the DWTP samples were PFOA and perfluorobutyric acid (PFBA) [44]. A similar study by Boiteux et al. investigated the presence of various PFAS in surface water and groundwater in France [47]. The investigation detected high PFNA, PFHxA, and PFOS concentrations at 52, 139, and 62 ng·L−1, respectively, in surface water, which according to the group could have resulted from increased industrial discharges of these compounds into rivers. Groundwater samples were high in concentrations of PFOA, PFBA, PFHpA, PFBS, PFHxS, and PFPeA, which were suggested to have resulted mainly from the breakdown of PFC precursors [47].
A study by Llorca et al. assessed the existence of PFASs in 127 water samples from Germany to Spain [50]. The study reported perfluorobutanoic acid (PFBA) as the more commonly found PFA in tap water, as it was in 54% of the samples. PFOS was the second most frequently found compound. Only six samples of the analysed drinking water presented concentrations of PFOS exceeding the threshold regulatory value [50]. Sun et al. investigated the presence of 10 PFASs in various matrices in Shenyang, China [55]. The group reported that PFC concentrations were higher in the Hun River (HR) after passing through two cities (Fushun and Shenyang). This indicates that cities contribute to PFC concentrations detected in waterways. In addition, an examination of the mass movement in the HR showed that Shenyang City (3.65 kg) contributed more PFOSs per year than Fushan City (1.29 kg) [55]. Overall, the concentrations of PFAS detected in Africa were higher compared to other countries, except for the 4,385 ng·L−1 concentration of PFOA detected in the Ruhr and Moehne rivers (Germany), as observed in Table 1. From Table 1, it could also be observed that the majority of the studies used solid-phase extraction (SPE) for enriching the target analytes prior to chromatographic detection. Only Flores et al. used liquid–liquid extraction (LLE) [53]. While in another study, dispersive SPE (d-SPE) was employed using graphitized carbon as the solid-phase sorbent [51]. Among the studies that used the SPE technique for enriching PFAS, the weak anion exchange (WAX) cartridge was the most used solid-phase material, except for a few that used hydrophilic–lipophilic balance (HLB) [46,55] and graphitized carbon [48,55] cartridges. The majority of the studies used the high-performance liquid chromatograph-tandem mass spectrometer (HPLC-MS/MS) for the detection of PFAS [41–45,47,49,50,53,54]. While Fagbayigbo et al. used UPLC-QTOF-MS [46], and others used HPLC or UPLC coupled with a triple quadrupole (QqQ) mass spectrometer operated in electrospray negative ionization mode (ESI-MS/MS) [48,51,52,55].
3 Health effects on humans, plants, and animals
Although they improve our quality of life and economically benefit many industries, PFASs pose a severe health threat to animals, plants, and humans [57]. Epidemiological studies indicated that PFOA and PFOS in humans contributed to kidney cancer, thyroid disease, ulcerative colitis, testicular cancer, pregnancy-induced hypertension, and high cholesterol [58–61]. Other effects include the disturbance of lipid homeostasis in the blood [62] and the alteration of biotransformation enzymes [63]. A study by Wang et al. on cat serum showed hyperthyroidism problems due to PFOS (15.8 ng·mL−1) exposure [64]. Another research on Daphnia Magna and Moina macrocopa indicated that PFOS (199.51 mg·L−1) and PFOA (17.95 mg·L−1) affected the performance of offspring [65]. The study also highlighted that PFOS was 10 times more toxic than PFOA [65]. In another study, the exposure of mice to sodium ρ-perfluorous nonenoxybenzene sulfonate (OBS) caused gut barrier dysfunction and hepatic metabolism disorder [66]. Research by Christie et al. studied the connection between size, sex, and the accumulation of PFAS in crocodiles in Kruger National Park (South Africa) [67]. No relationship was observed between exposure and crocodile size or sex. Geometrical differences were examined, and there was a significant difference between samples, depending on their sampling location. For example, the wet mass concentration of PFOS at Flag Boshielo Dam was 50.3 ng·g−1, which was higher than the other sampling sites with an average PFOS wet mass concentration below 14.0 ng·g−1, which suggested a point source of PFOS in this area [67]. A similar study on alligators also suggested no connection between the concentrations of PFAS and the sex and length of alligators [68]. PFAS concentrations varied across sampling sites. Alligators collected from Berkley County presented the highest PFOS concentrations (16.0 ng·g−1) and the overall number of detected PFAAs, which suggested that Berckley County was a point source of PFAS contamination [68].
3.1 ERA
ERA evaluates the hazard potential of existing or new environmental chemicals in the ecosystem [69]. Environmental exposure pathways include direct contact with soil, sediment, surface water, and eating contaminated food. Developing toxicity standards and choosing exposure assumptions are important when assessing the risk. For non-cancerous toxic effects, the risk process assumes that the worst outcomes occur only when the dosage (i.e., exposure) surpasses a regulatory value for a certain period [70]. To protect the health of human beings, regulatory agencies use this assumption to set threshold concentrations for chemicals. Figure 4 shows other key points when developing threshold values for chemicals in drinking water [71]. Globally, the regulatory concentration value for PFAS varies between agencies due to various methods for forming the threshold value, selecting different final data sets and the multiple frameworks among regional regulators [72,73]. The difference in the threshold value for PFAS indicates that regulatory bodies worldwide do not communicate with each other to come to a consensus on what the concentration limit for each or combined PFAS should be.
![Figure 4
Key points considered when developing threshold values for chemicals in drinking water, reproduced with permission from the study by Zodrow71et].](/document/doi/10.1515/revac-2023-0064/asset/graphic/j_revac-2023-0064_fig_004.jpg)
Key points considered when developing threshold values for chemicals in drinking water, reproduced with permission from the study by Zodrow71et].
Because of the limited or lack of data on the toxicity of many PFASs and the limited studies reporting on their adverse effects in humans, it is challenging to develop an ERA for PFAS. [74]. Most ecotoxicological investigations have fixated mainly on PFOS, PFOA, and a small variety of organisms (fish and water invertebrates), focusing on single‐chemical exposure [74]. Ecotoxicological research has also focused on testing organisms and their endpoints (development, survival, and reproduction) in water [74]. PFOS generally appears to have the highest relative toxicity among the PFAS studied [75]. However, this cannot be conclusive because many PFASs have not yet been reviewed. Hence, there are very little available data. Presently, toxicity data on a few PFAS, including PFOA and PFOS, have been used to develop screening standards for aqueous environments. For example, the federal water quality threshold of 6.8 μg·L−1 for PFOS set by Environment and Climate Change Canada (2017) [76] and 70 ng·L−1 for individual/combined concentrations of PFOA and PFOS by the USEPA [56]. Currently, there are no federally approved screening levels for ecological receptors. However, the USEPA has released draft water quality standards (USEPA REF#2300 and REF#2302) for PFOA and PFOS for civic comment [30].
Toxicity examinations on laboratory mammals (mice) have revealed that contact with PFAS may damage the hepatic, immune, and endocrine systems; trigger developmental complications; and certain kinds of cancers [30]. Despite studies showing that some human health problems can be associated with exposure to long-chain PFAS, current PFAS guidelines and toxicity influences (reference doses, cancer slope factors) are based on animal toxicity data [77]. The reliance on animal data is because there is a problem of concurrent exposure to multiple PFAS in the human study population, making it hard to examine the effect of a single PFAS. Despite this, the California EPA (2021) and USEPA established draft reference dosages (USEPA 2021 Ref#2258 and Ref#2259) for PFOA and PFOS and a draft cancer slope factor for PFOA (USEPA 2021 Ref#2258) using data obtained from human studies [30].
Proposed approaches to address serious gaps and doubts relating to ecological risk assessment for PFAS include [72]:
Development of targeted monitoring plans to support exposure evaluation.
Formulation of predictive models for bioaccumulation.
Development of computer models and laboratory/field methods that can competently evaluate biological effects and identify sensitive endpoints.
Encouragement of cross‐disciplinary approaches that use conventional and new procedures in a combined, resource‐effective way to address needs related to understanding the ecological risks of PFAS exposure.
Development of flexible tools that will cater to the special scale of contamination, which might range from localized (e.g., landfills) to global contamination (e.g., atmospheric transport).
Consideration of the large number of PFASs more than 4,500 chemicals, the importance of chemical mixtures, which may include precursors and degradation products, in developing approaches to estimating risk.
4 Analytical methods for determination of PFAS in the environmental matrices
Determination of PFAS in the environment is carried out using various analytical methods. For example, volatile PFASs in solids are typically analysed using gas chromatography and mass spectrometry (GC-MS) [78,79]. Ultrahigh-performance liquid chromatography coupled to high-resolution mass spectrometry (UHPLC-HRMS) (Q-orbitrap) has been used for biological samples such as liver and egg [80,81]. These samples require special care to evade the intrusion of endogenic compounds such as taurodeoxycholic acid that co-elute with PFOS and share similar MS/MS transitions [80,81]. Tandem mass spectrometry (MS/MS) with QqQ, a low-resolution detection method, is typically used for the analysis of PFAS in food/biological samples [82]. Combustion ion chromatography quantifies PFAS by measuring the total organofluorine (TOF), which captures all known and unknown PFAS, their precursors, and decomposition products without selectivity [83,84].
Although these techniques can determine PFAS in various matrices, most lack selectivity and sensitivity. Hence, sample preparation is required before detection to minimize contamination and maximize PFAS recoveries. Sample preparation consists of a series of steps called workflow [85]. The steps include extraction to remove or separate analytes from the aqueous media, pre-concentration to improve the obtained concentrations at small volumes, and clean-up to ensure high sensitivity and eliminate impurities that could interfere with chromatographic readings [86]. For example, the accepted standard workflow for evaluating PFOS in food samples includes solvent extraction, SPE, followed by liquid chromatograph-tandem mass spectrometer (LC-MS/MS) [87]. However, the workflow for PFAS in soil consists of separation using methanol and acetonitrile (ACN) followed by clean-up via SPE through WAX, HLB, and ENVI-Carb cartridges filled with activated carbon (AC) and graphitized carbon [88]. For evaluating PFAS in drinking water, the USEPA established an SPE-LC-MS/MS method (Method 537, Rev 1) which could analyse 14 different PFASs [89]. In 2018, the method was updated to include four more PFASs (Method 537.1). In November 2019, the technique was modified (Method 533) to analyse 29 PFASs, concentrating on PFAS with 4–12 carbon atoms [89]. Studying organic contaminants such as PFAS at trace levels in water matrices is often challenging as they cause poor reproducibility and analytical accuracy [90]. Sample preparation techniques typically used for water samples include SPE [91], LLE [92], ion-pair extraction [93], solid-phase microextraction (SPME) [94], and dispersive liquid–liquid microextraction (DLLME) [95]. With the constant appearance of new PFAS, the need for exceptional clean-up and analysis procedures also increases. The techniques are briefly explained in the subsequent sections, emphasizing their advantages and disadvantages.
4.1 SPE
4.1.1 SPE
SPE principle is based on separating the contaminant between a solid phase and an aqueous phase [96]. The SPE technique offers the advantage of adaptability with chromatographic analysis. It provides separation, concentration, decontamination, and clean-up [96]. The disadvantage of the SPE technique is that in water containing many foreign particles, for example, wastewater, the PFAS may bind to the particles [56]. Such water matrices require filtering of the water sample before SPE analysis to minimize low analyte recoveries caused by compound loss during the filtration step and the attachment of PFAS to the inside walls of the sample vessels [97]. Another limitation is the cartridge clogging when removing PFAS in heavily contaminated water samples. Conventional SPE-WAX cannot recover multiple PFAS in a single processing and measurement method. During SPE clean-up, neutral and positively charged PFAS are lost because of the high polarity range and the requirement for numerous separation techniques [97]. For example, van der Vegt et al. reported that using WAX-based SPE in their study did not permit the examination of PFOSA, PFTrDA, and PFTeDA at appropriate concentrations [97]. Still, the techniques proved to be an excellent clean-up method with low LOCs in multiple cases [97]. Excellent accuracies and relative standard deviation readings were also obtained. González-Barreiro et al. used SPE to preconcentrate 11 PFASs from wastewater [98]. The study reported that at lower pH (pH 4), they could only extract perfluoroalkyl carboxylates with a chain length of less than 10 carbon atoms. In contrast, at higher pH (pH 11), only five PFAS could be removed, namely, PFNA, PFOS, PFOA, PFDA, and PFOSA [98]. Silica-based adsorbants are mostly used in this technique due to their low cost, good extraction efficiency, and wide range of interactions with analytes (polar, nonpolar, and ion exchange) [99]. However, their lack of stability at very alkaline or acidic pH and difficulty in predicting their interaction mechanism with some analytes has led to exploring other solid materials [99]. For example, Deng et al. investigated using bamboo charcoal as an adsorbent for SPE separation of select PFAAs in various water systems [100]. The study reported a low detection limit of 0.01–1.15 ng·L−1, chemical stability, good reproducibility, and good repeatability. The bamboo charcoal had a surface area of 31.9 m2·g−1 [100]. Although charcoal adsorbents offer solutions to most of the problems of using SPE with other sorbents, the technique still suffers from a lack of selectivity because adsorbents like charcoal are attractive to a wide range of contaminants other than PFAS in water [99].
Various studies have applied the SPE technique and coupled it with analytical detection methods for analysing different PFAS across multiple water matrices, as shown in Table 2 [46,100–109]. The depicted data show that PFHpA, PFHxS, PFOA, PFNA, PFOS, and PFDA were the most investigated PFAS. Different PFAS were detected in various water matrices, including tap water, surface water, river water, barrelled drinking water, port water, sewage water, leachate, wastewater, seawater, and pond water. The most used detection method in SPE investigations was the HPLC-MS/MS, compared to UPLC-QTOF-MS and GC-ECD. In one of the studies, linearity was as narrow as 0.5–500 ng·L−1 [103], while another study had a wider linear range at higher concentrations between 5,050 and 286,400 ng·L−1 [105]. The limit of detection (LOD) was generally low (0.11–1.2 ng·L−1) when analysing PFAS from barrelled drinking, pond, port, tap, rain, and surface water [46,100,101,104], and was high (620–1,380 ng·L−1) in sewage water [105]. Similarly, the LOQ was lower (within 0.03–10 ng·L−1) for analysing PFAS in barrelled drinking, pond, port, tap, rain, river, and surface water [46,100,103,104], and higher (100–309 ng·L−1) in seawater analysis [107].
Analytical performance of traditional SPE coupled with various analytical detection methods for the analysis of PFAS in environmental matrices
| Analytes | Sample matrix | SPE sorbent | Detection technique | Linearity range (ng·L−1) | LOD (ng·L−1) | LOQ (ng·L−1) | Refs |
|---|---|---|---|---|---|---|---|
| PFPeA, PFHpA, PFOA, PFBS, PFNA, PFDA, PFUnDA, PFDoDA, PFOS | Surface water | HLB | UPLC-QTOF-MS | — | 0.02–0.08 | 0.065–0.261 | [46] |
| PFHpA; PFHxS, PFOA, PFNA, PFOS, and PFDA | Barrelled drinking water, port, tap, and pond water | Bamboo charcoal | LC-MS/MS | 0.1–1,000 | 0.01–1.15 | 0.03–3.85 | [100] |
| PFOA | Tap water and rainwater | Bamboo charcoal | LC-MS/MS | 1–1,000 | 0.2 | — | [101] |
| PFOA | Raw and treated leachate | WAX | LC-MS/MS | — | — | — | [102] |
| PFOA and PFOS | Distilled, tap, and river water | MWCNT-R-NH2 | UPLC-ESI-MS | 500–10,000 | — | 10 and 15 | [103] |
| PFHxS, PFHpA, PFNA, PFOS, PFOA, and PFDA | Barrelled drinking water, pond, and tap water | CNSs-COOH | LC-MS/MS | 0.50–200 | 0.01–1.2 | 0.03–4.0 | [104] |
| PFOA, PFPeA, and PFHpA | Sewage water | TR-5 capillary | GC-ECD | 5,050–286,400 | 620–1,380 | — | [105] |
| C5–C17 | Wastewater | C18 SPE column | HPLC-LTQ-Orbitrap-MS | — | — | — | [106] |
| PFPA, PFHxA, PFHpA, PFOA, PFNA, PFDA, PFDoDA, PFTrDA, PFBS, PFHxS, PFOS, PFOSA | Seawater | HLB and Strata™-X | HPLC-ESI-MS/MS | — | — | 100–309 | [107] |
| PFOA, PFOS, PFNA, PFHxS, PFBS, FOSA, EtFOSA, MeFOSA, EtFOSE, MeFOSE | Landfill leachate | WAX | LC-MS/MS | — | — | — | [108] |
| 30 PFAS | Drinking water | Styrene divinylbenzene (SDVB) | UPLC-QTOF-MS | — | — | — | [109] |
4.1.2 Dispersive solid phase extraction
d-SPE is a sample preparation technique focusing on the dispersion of a small amount of adsorbent in a sample solution through ultrasonication or vertexing [110]. Once analytes are extracted, centrifugation separates analyte-bound adsorbents from the solution, followed by elution of the analytes using small volumes of organic solvent [110]. The d-SPE method’s advantages are that it does not utilize a large mass of adsorbent, offers high analyte recoveries, uses a low volume of solvent, and uses maximum adsorbent surface area owing to the use of dispersion and sonication, which allow for maximum interaction between the sorbent material and the target analyte [111]. Compared to the traditional SPE technique, wherein the solid-phase material is stationary and fixed as part of the cartridge, the analyte solution is passed through and thus may not allow enough interaction between the sorbent and analyte long enough to achieve maximum sorption [111]. The adsorbent is added directly to the analyte solution to avoid loss of analytes due to any washing steps as observed in traditional SPE methods [110,111]. Columns do not clog compared to traditional SPE, since no column is required. Traditional SPE techniques use columns or cartridges as the sorbent material, whereas in d-SPE, the sorbent material is directly introduced into the sample solution. The challenges with d-SPE include difficulty in attaining comparable detection limits as a conventional SPE technique, reduced efficiency, restricted concentrations of the raw extract to a small end volume due to ionization effects, high dead volume, and peak tailing, as a result, of dissolved residual salts in the MS source (and LC column) [110,111].
Because of the strong binding characteristics of the d-SPE-salts, it is difficult to attain solid phases with zero PFAS trace. An initial washing step is included in conventional SPE procedures to eradicate most PFAS residues in the SPE column. This preconditioning phase is impossible for d-SPE, possibly causing higher background concentrations [110,111]. Another difficult part of d-SPE is that the quantity of the sorbent material depends more on the medium’s water content, making the technique less robust for varying environments. Few studies have applied this technique for PFAS analysis in water (Table 3). For example, Ma et al. developed a fluorinated-metal organic framework material (F-MOF) as a sorbent in d-SPE and adsorbed 419.8 mg·g−1 PFOA in various water matrices [112]. The F-MOF-based SPE was coupled with GC-MS analysis, and the LOD was 2.6 ng·L−1 [112]. In another study by van der Vegt et al., d-SPE and traditional SPE were used comparatively to analyse PFAS in apples [97]. The study reported that using d-SPE was easy, allowing non-ionic and ionic PFAS analysis concurrently [97]. The LOCs were commonly greater owing to (i) a lesser medium-to-column ratio and (ii) background signals on some PFAS. PFAS such as PFOS, PFOA, PFTeDa, PFTrDa, PFDoDa, PFUnDa, PFDA, and PFNA were detected in most studies conducted using the d-SPE technique. However, most studies did not report the linearity range [112,114–117,121,122]. The lowest LOD was obtained by Li et al. at 0.02–0.16 ng·L−1 using M-FPCs to analyse PFAS in the rain and lake water [114], and the highest LOD was 4–40 ng·L−1 as reported by Huang et al. investigating PFAS in tap and lake water using Fe3O4@CB [113]. The most used analytical technique was LC-MS/MS. Huang et al. reported the highest (5–100 ng·L−1) LOQ compared to Lin, Fan, and Selahle, and their co-workers, respectively, whose LOQ ranged between 0.06 and 2.2 ng·L−1 [114,116,118].
Analytical performance of d-SPE coupled with various analytical detection methods for analysing PFAS in environmental matrices
| Analytes | Sample matrix | d-SPE sorbent | Detection technique | Linearity (ng·L−1) | LOD (ng·L−1) | LOQ (ng·L−1) | Refs |
|---|---|---|---|---|---|---|---|
| PFOA | Tap, lake, and bottled drinking water | F-MOF | GC-MS | — | 2.6 | 8.7 | [112] |
| PFUdA, FHEA, 6:2 FTS, PFOS, PFOA, PFHxS, PFPeS, HFPO-DA, and 4:2 FTS | Tap and lake water | Fe3O4@CB | UHPLC-HRMS | 10–10,000 | 4–40 | 5–100 | [113] |
| PFOS, PFOA, PFTeDa, PFTrDa, PFDoDa, PFUnDa, PFDA, and PFNA | Rain and lake water | M-FPCs | UPLC-MS | — | 0.02–0.16 | 0.06–0.53 | [114] |
| PFHpA, PFHxS, PFNA, PFOS, PFDA, PFOA, PFBA, PFDoA, and PFUdA | Drinking water | ZIF-67/g-C3N4 | LC-MS/MS | — | 0.3–2 | — | [115] |
| PFOA, PFDA, PFNA, PFUnDa, PFDoDa, PFOS, and PFTeDa | Wastewater | F-COPs | UPLC-MS | — | 0.05–0.13 | 0.17–0.43 | [116] |
| PFBS, PFOS, PFOA, PFBA, PFPeA, PFHxA, and PFHxS | Reagent water | Graphitized carbon black | LC-ESI-MS/MS | — | — | — | [117] |
| PFOA, PFDA, PFOS, and PFTeDA | River water | Fe3O4@MIL-101(Cr) | UHPLC-MS/MS and HPLC-DAD | 1–5,000 and 0.05–2,000 | 0.3–0.66 and 0.011–0.04 | 1.0–2.2 | [118] |
| 23 PFAS | Water | C18 sorbent and graphitized carbon black. | UPLC-MS/MS | — | 0.5–10 | 2–20 | [119] |
| PCAs | Water | TPTGCl | LC-MS/MS | 3–625 | 0.43–0.78 | — | [120] |
| PFOS | Water and wastewater | Chitosan-MIP | HPLC | — | — | — | [121] |
| PFNA, PFBS, PFHpA, PFHxS, PFOS, PFHxA, PFDA, and PFPeA | Tap water | MWCNTs@MIPs | HPLC | — | — | — | [122] |
4.1.3 Magnetic solid-phase extraction (MSPE)
A new SPE mode dubbed MSPE has recently been established. It has numerous benefits in parallel with conventional SPE techniques [123,124]. After use, the procedure uses a magnet to separate the adsorbent–analyte complex [125]. Thus, there is no need to centrifuge, making the recovery of the sorbent material faster and easier [123,124]. Various studies that conducted the MSPE technique are summarized in Table 4. In an investigation by Xian et al., 19 PFAS were separated from different water mediums (tap, river, and drinking) using novel fluorine and nitrogen-functionalized magnetic graphene (G-NH-FBC/Fe2O3) [125]. The study achieved linearity within 500–200,000 ng·L−1, with LOD and LOQ ranges of 3–15 and 10–49 ng·L−1, respectively. The removal efficiency range was 71.9–117.6% [125]. A similar study used magnetic decorated graphene sheets (MG@CTAB/SDS) to recover PFDoA, PFOS, PFDA, PFTA, and PFOA from river water and effluents [126]. The obtained percentages were between 56.3% and 93.9% [126]. Ren et al. used covalent organic frameworks (COFs) functionalized with magnetic nanoparticles to remove PFAS from water matrices [127]. The group reported LOD and LOQ ranges of 0.62–1.39 and 2.16–4.63 ng·L−1, respectively [127]. The successful preparation of a one-pot synthesized magnetic fluorinated carbon nanotube adsorbent (MFCA) allowed for extracting multiple PFAS from water [128]. The study achieved detection limits of 0.010–0.036 and 0.024–0.50 ng·L−1 for PFCAs and PFSAs, respectively [128]. Hydrophobic, fluorous–fluorous, and hydrogen bonding interactions were responsible for the removal of the PFAS by MFCA [128]. In another study by Song et al., an adsorbent based on the functionalization of multiwalled carbon nanotubes with magnetic carbon (MWCNTs@Fe3O4@C) achieved good repeatability (3.80–9.52%) and low LOD (0.03–0.09 ng·L−1) [129].
Analytical performance of MSPE coupled with various analytical detection methods for the analysis of PFAS in environmental matrices
| Analyte | Sample matrix | Adsorbent | Detection method | Linearity (ng·L−1) | LOD (ng·L−1) | LOQ (ng·L−1) | Ref |
|---|---|---|---|---|---|---|---|
| PFDoA, PFNA, and PFOA | Lake water | F17-Fe3O4@mSiO2 | GC-MS | 200–200,000 | 55–86 | 180–280 | [124] |
| 19 PFAS | Drinking water, river water, tap water, factory drainage | G-NH-FBC/Fe2O3 | LC-HRMS | 500–200,000 | 3–15 | 10–49 | [125] |
| PFDoA, PFOS, PFDA, PFTA, and PFOA | River water and effluents | MG@CTAB/SDS | HPLC-ESI-MS/MS | 1–500 | 0.15–0.50 | — | [126] |
| PFBA, PFPeA, PFHpA, PFOA, PFNA, and PFDA | Purified water, river, snow, and pond water | CTF/Fe2O3 | LC-MS/MS | 5–4,000 | 0.62–1.39 | 2.16–4.63 | [127] |
| PFBS, PFHXS, PFHPS, PFOS, PFDS, PFNA, PFHA, PFOA, PFDA | Lake water, river water, tap water, and wastewater | MFCA | LC-MS/MS | 1.0–10,000 | 0.010–0.50 | 0.034–1.66 | [128] |
| PFHxS, PFOA, PFHpA, PFNA, PFDA, and PFOS | Tap, snow, and barrelled drinking water | MWCNTs@Fe3O4@C | LC-MS/MS | 0.1–1,000 | 0.03–0.09 | 0.1–03 | [129] |
| PFDoA, PFHpA, PFDA, PFTeDA, PFOA, and PFOS | River water | Fe3O4@SiO2@FBC MNPs | UHPLC-MS/MS | 0.25–25 | 0.01–0.06 | 0.05–0.20 | [130] |
| PFOA, PFOS, PFNA, PFDA, PFUnDA, FDoDA, and PFTA | Wastewater | Fe3O4-C18-chitosan MNPs | HPLC-ESI-MS/MS | 0.5–50 | 0.033–0.19 | — | [131] |
| PFOS, PFTA, PFOA, PFNA, PFDA, PFHxS, PFHpA, PFDoDA, and PFUnDA | Surface water | Fe3O4@SiO2-NH2&F13 | UPLC-MS/MS | 0.5–100 | 0.029–0.099 | 0.097–0.330 | [132] |
| PFNA, PFDA, PFDoA, and PFUdA | Drinking water, groundwater, and lake water | Fe3O4@EB-iCOFs | LC-MS/MS | 1–1,000 | 0.1–0.8 | 0.3–2.5 | [133] |
By using the MSPE technique, PFHxS, PFOA, PFHpA, PFNA, PFDA, and PFOS were the most commonly detected PFAS in water (Table 4) [124–133]. Of most importance is that PFASs were detected in a wide range of water matrices, including lake water, river water, tap water, pond water, drinking water, snow, effluents, groundwater, surface, and wastewater [124–133]. The lowest LOD was 0.01 ng·L−1 using MFCA [128] and Fe3O4@SiO2@FBC MNPs as adsorbents [130]. The lowest LOQ was 0.034 ng·L−1 using MFCA [128]. The most used analytical detection method was LC-MS/MS.
4.1.4 SPME
SPME is a method that combines sampling, pre-concentration, and clean-up in a single step [134]. Various SPME techniques exist, such as headspace SPME (HS-SPME), multiple monolithic fibre SPME (MMF-SPME), and online SPME. Compared to conventional extraction techniques, SPME allows for the separation of organic compounds from both biological and aqueous environments at low sample volumes. It offers advantages such as reduced preparation time, disposal costs, and solvent use. SPME provides lower LOD when separating volatile and semi-volatile organic compounds from ecological samples [135]. In a study by Saito et al., an online in-tube SPME technique was developed and coupled with LC-MS to analyse PFOA and PFOS in the river and tap water samples [136]. The method offered convenience for automating the separation process, decreased analysis time, and enhanced sensitivity and precision compared to manual extraction. The LOD for PFOS and PFOA were 1.5 and 3.2 ng·L−1, respectively [136].
A HS-SPME is an extraction procedure that traps and concentrates analytes from a dynamic or static headspace process using a coated fused silica fibre [137]. For example, Bach et al. investigated selected volatile PFASs in sediment and water samples using a validated HS-SPME technique coupled with GC-MS. The authors obtained recoveries between 76 and 126% for water samples, which varied per PFAS type and water sample type, and the LOQ range was 20–100 ng·L−1 [137]. HS-SPME is reported to offer rapid analysis and simplicity and does not utilize any solvent. With reduced preparation time, the technique allows the study of multiple samples simultaneously in different sample matrices [138,139]. Another technique is the numerous MMF-SPME, enabling PFCAs from complex water matrices to be extracted efficiently [140]. Huang et al. used MMF-SPME to analyse tap water, wastewater, milk, and river water for PFCAs [141]. The authors reported that the technique was cost-effective and simple to operate and had a fast mass transfer. Their LOD ranged from 0.40 to 12.1 ng·L−1 [141].
In a study by Lockwood et al., µSPE cartridges were filled with mixed-mode C18: aminopropyl silica as the solid phase for extracting 13 short-chain and long-chain PFASs from surface water samples [142]. The results were similar to those obtained using standard SPEs but required a smaller sample volume of about 2 mL, and the preparation time was reduced to 5 min [142]. The study recorded recoveries between 86 and 111%, with the concentrations of PFAS at nearly 900 ng·L−1 in the tested surface water sample [142].
By using MSPE, the most detected PFAS in water were PFBA, PFPA, PFHA, PFOA, PFNA, PFDA, and PFOS (Table 5) [136–147]. Fibre MSPE and online SPME were the most used the type of MSPE. As observed in Table 5, tap water, wastewater, and river water were the most analysed water matrices. Linearity was generally high for most studies, and the range was wide. The lowest range was 1–50 ng·L−1 obtained by Lashgari et al. [146]. The lowest LOD range was 0.02–0.8 ng·L−1 using COF-F-1 and MCM 41 [143,146]. The LODs obtained were generally low compared to other sample preparation techniques. The lowest LOQs were 0.06–3 ng·L−1 and 0.07–0.28 ng·L−1. Bach et al. reported the highest LOQ range of 898 ± 15 ng·L−1 [137]. Generally, the LOQ values were more than 1 ng·L−1 for most studies. The LC-MS/MS technique was the most used detection method.
Analytical performance of SPME coupled with various analytical detection methods for the analysis of PFAS in environmental matrices
| Analytes | Sample matrix | Adsorbent | SPME type | Detection technique | Linearity (ng·L−1) | LOD (ng·L−1) | LOQ (ng·L−1) | Refs |
|---|---|---|---|---|---|---|---|---|
| PFOA and PFOS | Tap water, river water, pond water, and water eluate | GC capillary column | Online-SPME | HPLC-MS | 5–5,000 | 1.5–3.2 | 5–11 | [136] |
| Volatile PFASs | Tap water, surface water, and sediment | Silica fibre | HS-SPME | GC-MS | — | — | 20–100 | [137] |
| PFBA, PFPA, PFHA, PFOA, PFNA, PFDA | Tap water, wastewater, and river water | DFHA and VTAC | MMF-SPME | LC-MS/MS | 2.5–150,000 | 0.40–12.1 | 1.32–14.5 | [141] |
| PFOS, PFNS, PFHpS, PFDS, PFHpA, PFOA, PFNA, PFDA, PFUnDA, PFHxDA, PFODA, PFDoDA, PFTrDA, and PFTeDA | Pure water | COF-F-1 | SPME | NanoESI-MS | 1–5,000 | 0.02–0.8 | 0.06–3 | [143] |
| PFOS and PFOA | Surface water | TiO2 NTs/Ti | Fibre | HPLC-MS | 10–10,000 | 2.5–7.5 | — | [144] |
| Polybrominated diphenyl ethers (PBDEs) | Tap water, river water, and wastewater | Bamboo charcoal/Fe3O4 | Fibre | GC-NCI-MS | 1–1,000 | 0.25–0.62 | — | [145] |
| PFDA, PFNA, PFBA, PFPeA, PFUnA, PFHpA, PFHxA, PFOA, PFDoA, PFTeDA, PFTeDA, PFHxS, PFOS, PFBS | Surface water | C18: aminopropyl silica | — | LC-MS/MS | — | 0.29–6.6 | 898 ± 15 | [142] |
| PFPA, PFHpA, PFOA, PFNA, PFDA | River water and rainwater | MCM-41 | — | LC-MS/MS | 1–50 | 0.02–0.08 | 0.07–0.28 | [146] |
| PFDoA, PFHxS, PFOA, PFNA, PFHpA, PFDA, PFOS, PFUnA | River water, wastewater, and seawater | NH2-ZIF-8 | Fibre | LC-MS/MS | 1–5,000 | 0.15–0.75 | — | [147] |
4.2 Liquid extraction
Liquid extraction is a conventional sample preparation procedure that allows the separation and pre-concentration of analytes into smaller volumes of the extraction solvent using a triple solvent system [148]. The working principle of this technique involves the injection of a small volume of a suitable separation solvent and a disperser solvent into the aqueous solution [148,149]. Several strategies for LLE, such as traditional LLE, DLLME, vortex-assisted liquid–liquid microextraction (VALLME), and BHF-LPME, have been proposed and applied for the analysis of PFAS in water [98,149–163]. VALLME is a technique that was developed by Yiantzi et al., wherein a vortex stirrer mechanically stirs a suspension achieved by LLME to disperse a low-density extractant organic solvent into the aqueous phase [152]. This results in the formation of tiny droplets containing the analytes that can be centrifuged to decant the two phases and separate them [153]. The advantages of this technique include low detection limits, fast analyte extractions, good repeatability and reproducibility, simplicity, high sensitivity, improved extraction efficiency, and no sample matrix effect [152,154]. The major disadvantages of this technique include the limited number of suitable extractants and the drop stability, which can be impaired by high stirring speed, long extraction time, high temperature, and solid particles in suspension [155,161]. Table 6 summarises studies applying liquid extraction techniques to detect PFAS in various water matrices, including tap water, river water, seawater, wastewater, canal water, well water, and seawater [98,150–163]. Tap and river water were the most analysed matrices among the different water samples. The most commonly detected PFASs using liquid extraction were PFOS and PFOA. Various extraction solvents were used, namely, perfluoro-tert-butanol, octanol, MTBE, undecanol, chloroform, dibutyl ether, butanol, hexanol, hexane, and ACN [98,150–163]. Amongst them, octanol was the most used by many studies. HPLC-MS/MS was the most used detection method amongst others that were used, such as GC-MS/MS, HPLC-ESI-LTQ-HRMS, HPLC-ESI-MS/MS, and LC-QToF [98,150–163]. The linearity range was lowest by 0.5–300 ng·L−1 and high by 1,000–2,000,000 ng·L−1 [156,163]. The LODs and LOQs were generally low for most studies except for a few, where the range exceeded 100 ng·L−1. Wide linearity, LOD, and LOQ ranges (10–1,000,000, 1–1,446, and 3–1,080 ng·L−1) were obtained when using undecanol as the extraction solvent and DLLME-SFO as the sample preparation technique [157].
Analytical performance of liquid phase-based methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Analytes | Sample matrix | Technology | Extractant | Detection technique | Linearity (ng·L−1) | LOD | LOQ | Refs |
|---|---|---|---|---|---|---|---|---|
| 20 PFAS | Tap and river water | DLLME | Perfluoro-tert-butanol | LC-MS/MS | 0.5–500 | 0.19–4.2 | 0.6–8.7 | [150] |
| PFOS | River, well, and tap water | VALLME | Octanol | LC-MS/MS | 5–500 | 1.6 | 5 | [154] |
| MeFOSA and EtFOSA | Seawater | VALLME | Octanol | HPLC-ESI-LTQ-HRMS | 0.5–300 | 0.22–3.0 | 0.7–6 | [156] |
| PFDoA, PFHxA, PFOA, PFNA, PFHpA, PFDA, PFOSA, PFUnA, PFDS, and N-EtPFOSA | Wastewater | LLE | MTBE | HPLC-ESI-MS/MS | — | 0.26–0.62 | 0.94–2.3 | [98] |
| PFOS, PFOA, PFHpA, PFHxA, PFPeA, and PFBuA | Tap and surface water | DLLME-SFO | Undecanol | LC-MS/MS | 10–1,000,000 | 1–1,446 | 3–1,080 | [157] |
| PFHA, PFOA, PFDA, and PFNA | Seawater and river water | DLLME | Chloroform | GC-MS/MS | 250–2,000,000 | 37–51 | 123–170 | [158] |
| PFHpA, PFOA, PFDoA, PFUdA, PFTrDA, PFTeDa, and PFNA | Canal water | BHF-LPME | Dibutyl ether | UHPLC-MS/MS | 5–10,000 | 0.40–6.48 | 1.25–224 | [159] |
| PFOA | Deionized and tap water | LLE | Butanol, hexanol, and octanol | LC-MS/MS | — | — | — | [160] |
| PFOA | Water | LLE | Octanol | LC-MS/MS | — | — | — | [162] |
| PFAS | Water | — | Hexane, MTBE, hexane: MTBE (1:1/v/v), and ACN | LC-QToF | 1,000–2,000,000 | — | — | [163] |
5 Removal of PFAS from water
Traditional treatment techniques such as disinfection of PFAS by free chlorine, UV irradiation, and oxidation using ozone and hydroxyl radicals are ineffective for eradicating PFAS from water [164,165]. This is because of the strong C-F interaction, their functional groups’ electron-extracting property, low water concentration, and high hydrophilicity [165]. Biological treatments, both aerobic and anaerobic, can only disrupt the C-C interaction to form short-chain PFAS.
5.1 Chemical oxidation
5.1.1 Heat-activated persulfate
Most studies focus on activated persulfate (
Analytical performance of heat-activated persulfate-based chemical oxidation methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology | Sample matrix | Target analyte | By-products/intermediate | Degradation efficiency (%) | Linearity (ng·L−1) | Detection method | Refs |
|---|---|---|---|---|---|---|---|
| Heat-activated persulfate | Modelled ultra-pure water | PFOA, PFHpA, PFHxA, PFPeA, PFBA, PFPrA | C2–C6 | 90 | — | 881 Compact IC pro-anion | [166] |
| Heat-activated persulfate | Modelled ultra-pure water | PFOA, 6:2 fluorotelomer sulfonate, PFOS | C4–C7 | >90, No degradation, 98 | — | HPLC-ESI/MS/MS | [167] |
| Heat-activated persulfate | Groundwater | PFOA | C2–C7 | 89.9 | — | LC-MS/MS | [168] |
| Heat-activated persulfate | Modelled ultra-pure water | PFOS | C0–C6 PFCAs | — | — | LC-MS/MS | [170] |
| Heat-activated persulfate | Modelled ultra-pure water | PFOA | C4–C7 PFCAs and F− | 98 | — | LC-MS/MS | [169] |
| Heat-activated persulfate | Tap water, groundwater, and modelled ultra-pure water | PFOA | C4–C7 PFCAs, CO2 and F− | 85 | — | LC-MS | [171] |
| Heat-activated persulfate | Modelled ultra-pure water | 6:2 FTS, PFOA, and PFOS | C3–C7 | 46.5–100 | — | LC-MS/MS | [172] |
| Heat-activated persulfate | Groundwater | PFOA | C2–C6 | 99.8 | 1,000–1,000,000 | HPLC-MS | [173] |
| Heat-activated persulfate | Modelled ultra-pure water | PFNA, PFOA, and PFHpA | C4–C8 | 30–83 | — | HPLC-MS | [174] |
| Heat-activated persulfate | Modelled ultra-pure water | PFNA, PFOS, PFHpA, PFHxS, PFBS, PFOA and PFOS | C4–C8 | — | — | HPLC-MS | [175] |
5.1.2 Electrochemical oxidation
Electrochemical oxidation occurs through direct and indirect electrolysis. Direct electrolysis involves the adsorption and degradation of contaminants directly at the electrode. In contrast, indirect electrolysis consists of the degradation of pollutants in the bulk liquid in reactions with oxidizing agents developed at the electrode [176]. Electrodes effective for the use of PFAS include titanium oxide (TiO2), lead dioxide (PbO2), tin oxide (SnO2), and boron-doped diamond (BDD). Electrochemical oxidation has numerous benefits over other oxidation techniques, such as operating at room temperature and not requiring a chemical [177]. The downside of this technique is the production of by-products such as hydrogen fluoride (HF), bromate, lead, and perchlorate, which are toxic, expensive electrode costs, and secondary pollution due to possible release into the environment [177]. As a solution to the problem of perchlorate production as a by-product during the electrochemical breakdown of PFAS, Schaefer et al. introduced a biological treatment polishing step [178]. Many studies have been conducted to remove various PFAS, including PFOA, PFOS, PFHxS, 6:2 FTAB, 6:2 FTSA, and other PFCAs (C3–C18) from water matrices using electrochemical oxidation-based technology [177–185]. Amongst the different electrodes that can be used for PFAS oxidation, most studies reported using BDD, and only a few used a combination of two or more (Table 8). Electrochemical oxidation studies were conducted in a small variety of water matrices, which included modelled ultra-pure water, groundwater, wastewater, distilled, and river water [177–185]. Amongst the matrices, modelled ultra-pure water was the most prevalent. The variety in terms of the types of PFAS investigated was also narrow, seeing that only a single study by Barisci et al. investigated a wide range of PFCAs with carbon chains between C3 and C18, including PFOA [185]. At the same time, most studies targeted PFOA [177–180,183,184]. Only a few studies included PFOS as a target analyte [178–180]. The results clearly show the biasedness of most electrochemical oxidative studies towards PFOA. Generally, the electrochemical oxidation of PFAS was satisfactory, with percentage degradation efficiencies of more than 80% for most of the studies. In a comparative study by Gomez-Ruiz et al., microcrystalline (MCD)/BDD and ultra nanocrystalline (UNCD/BDD) were investigated for electrochemical oxidation of PFOA in modelled ultrapure-water. The study reported the highest (100%) and the lowest (21%) of the obtained degradation efficiencies. Degradation of PFOA and other PFASs resulted in various by-products or intermediates, including F−, Cl−, TBA, PFCAs (C2–C8),
Analytical performance of electrochemical oxidation-based methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Electrode | Sample matrix | Target analyte | By-products/intermediates | Degradation efficiency (%) | Detection method | Refs |
|---|---|---|---|---|---|---|
| Ti/SnO2-Sb2O5/PbO2 | Modelled ultra-pure water | PFOA | C6–C7 and F− | 94 | GC-MS | [177] |
| BDD and Ti | Modelled ultra-pure water | PFOS, PFHxS, and PFOA | — | 88.1–94 | LC-MS/MS | [179] |
| BDD | Groundwater | PFOS and PFOA | Cl− and TBA | >60 | LC-MS/MS | [178] |
| Ti/RuO2-IrO2@Pt | — | PFOA | N/A | 99.30 | — | [180] |
| Ti/Sn-Sb/SnO2-F-Sb | Wastewater | PFOS | F−, PFCAs (C2–C8) | 99 | UPLC-MS/MS | [181] |
| BDD | Wastewater | 6:2 FTAB, 6:2 FTSA M4 | C6–C7,
|
97.1 | UHPLC-MS/MS | [182] |
| Ti/SnO2-Sb2O5/PbO2-PVDF | Modelled ultra-pure water | PFOA | — | 92.1 | HPLC-UV | [183] |
| MCD/BDD ultra nanocrystalline (UNCD/BDD) | Modelled ultra-pure water | PFOA | F− | 100 and 21 | HPLC-MS/MS and HPLC-DAD | [184] |
| Si/BDD | Wastewater, distilled, and river water | PFOA, PFCAs (C3–C18) |
|
39–95 | LC-MS/MS | [185] |
5.1.3 Photolytic/photochemical oxidation
Photochemical oxidation is defined as a chemical reaction in which an organic or inorganic chemical is broken down by photons [186]. The disadvantages of photochemical oxidation are the generation of toxic and persistent intermediates, reduced reaction rates, limited catalyst stability, and difficult catalyst recovery [187]. The mineralization of pollutants is sometimes incomplete due to the formation of intermediates [187]. Photolysis occurs directly (photooxidation) or indirectly (photochemical oxidation). During photooxidation, pollutants are transformed or degraded through the absorption of UV radiation. In photochemical oxidation, reactive species interact with pollutants to alter or degrade them [188]. UV irradiation is divided into four regions: UVA (315–400 nm), UVB (280–315 nm), UVC (200–280 nm), and vacuum UV (VUV, 100–200 nm) [186]. Low-pressure (LP) mercury lamps emitting UVC irradiation primarily at 254 nm have been investigated for the degradation of various contaminants, including PFAS. While it is an efficient process for a range of compounds, conventional direct photolysis is known to be inefficient for the degradation of PFAS, as demonstrated by several investigations [189–191]. For example, Chen et al. investigated the degradation of PFOA using two types of LP mercury lamps emitting light at 254 and 185 nm [189]. PFOA was significantly decomposed under irradiation of 185 nm light, while it was very slow and negligible under 254 nm light irradiation. The observation was due to its strong absorption of PFOA from the deep UV region to 220 nm and a weak absorption from 220 to 460 nm [189]. Little to no degradation during direct UV photolysis is attributed to the insufficient breakdown of the C-F bond by photo energy generated during UV irradiation [186]. In an interesting study by Chen et al., hydrogen polarization was used to enhance degradation and defluorination of PFOA and PFOS in tap and simulated groundwater [192]. The group reported a degradation of 95% for PFOA and 87% PFOS using the H2-VUV system, with PFOA showing six short-lived short-chained intermediates that were completely degraded at the end of the reaction [192]. In contrast, H2-polarized VUV degradation of PFOS produced one H/F exchange intermediate (
Other similar studies are summarised in Table 9, including their performance [189–199]. Overall, most studies on photolytic degradation of PFAS used UV or VUV and analysed PFAS in modelled ultra-pure water. PFOA was the most targeted PFAS compared to TFHFESE, 6:2 FTS, PFNA, PFHpA, PFBA, PFOS, GenX, PFHxS, and PFOS [189–191,193–199]. During the photolytic degradation of PFAS, the most prevalent by-products or intermediates were C2–C7 PFCAs, F−, CO2, and
Analytical performance of photochemical degradation methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology | Sample matrix | Target analyte | By-products/intermediates | Degradation efficiency (%) | Detection method | Refs |
|---|---|---|---|---|---|---|
| VUV and UV | Modelled ultra-pure water | PFOA | C4–C7, F− | 61.7 | HPLC-MS | [189] |
| VUV and UV | Modelled ultra-pure water | PFOA | CH₂O₂, CH₃COOH | 100 | — | [190] |
| UV | Modelled ultra-pure water | PFOA | C3–C7, F−, and CO2 | 100 | HPLC-CDD | [191] |
| UV/H2O2 and tungstic heteropolyacid photocatalyst | Wastewater | PFOA | C4−C6, F−, CO2 | 100 | GC-MS | [193] |
| UV, UV/H2O2, solar, solar/H2O2, O3, and O3/H2O2 | River water | TFHFESE | F− | 98.7–100 | GC-MS | [194] |
| UV, UV/H2O2, O3, O3/H2O2, and Fe2+/H2O2 | Modelled ultra-pure water | 6:2 FTS | C2–C7,
|
97 | HPLC-CDD | [195] |
| Sulfite/UVC and TiO2/UVC | Modelled ultra-pure water | PFNA, PFOA, PFHpA, PFBA, PFOS, GenX, PFHxS, 6:2 FTS, PFBS | C5–C8 | 95–100 | HPLC-MS | [196] |
| UV | Groundwater | PFOS | C2–C8, F− | 99 | HPLC-MS | [197] |
| UV/Iodide/HA | Modelled ultra-pure water | PFOS | C4, C6, and C8 | 86 | LC-MS/MS | [198] |
| UV/Iodide | — | PFOA | F−, C2–C6 | 100 | GC-MS | [199] |
| H2-polarized VUV | Tap water and simulated groundwater | PFOA and PFOS |
|
87–95 | UPLC-HRMS | [192] |
5.2 Ultrasonication (US)/sonolysis
Sonolysis uses ultrasonic irradiation without catalysts to produce hydroxyl radicals in aqueous matrices [200–202]. It is a successful system for degrading organic pollutants in aqueous media [200]. Sonolytic breakdown follows two mechanisms: (1) hydroxyl radical reaction because of the homolytic separation of water under severe conditions and (2) pyrolysis. The sonolytic process depends on disseminating acoustic waves in liquids at frequencies ranging between 20 and 1,000 kHz, which results in cavitation. Throughout cavitation, cyclic formation, growth, and collapse of microbubbles cause an intense rise in temperature (5,000 K), pressure (2,000 atm), and the generation of free radicals [201]. Chemical reactions occur at different active sites within the cavitating bubbles: (1) inside the bubble gas, (2) at the bubble gas–liquid interface, and (3) in the bulk liquid area. Pyrolysis happens more in the bubble gas, and hydrophobic compounds like PFOS and PFOA tend to accumulate at the bubble–water interface owing to their high surface activity. Factors that affect the efficiency of sonolysis include power density, frequency, sparge gas, temperature, and initial concentration of PFAS. According to Sidnell et al., a meta-analysis of PFAS showed that mid- to high-frequency (100–1,000 kHz) sonolysis mechanisms are similar, regardless of reaction conditions, and that PFAS degrades quickly (0.002–0.04 min−1) at mid frequency without the addition of oxidative agents, compared to low frequencies (0.0002–0.03 min−1) [203]. Various studies on the sonolytic degradation of PFAS and their performance are summarised in Table 10 [201–208]. Most of the studies were conducted at mid frequency (100–1,000 kHz), except for one investigation by Wood et al., which included a lower frequency of 44 kHz [207]. Only two types of water matrices were investigated for sonolytic degradation, modelled ultra-pure water, and groundwater. PFOS and PFOA were the main targets of PFAS in many of the studies, except for two studies in which one only investigated 6:2 FTAB and HFPO, and the other at least included a wider range of PFAS by investigating 15 different PFASs [205,206]. The main by-products or intermediates obtained for sonolytic degradation of PFASs included C2–C7, CO, CO2,
Analytical performance of sonolytic degradation methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology frequency (kHz) | Sample matrix | Target analyte | By-products/intermediates | Degradation efficiency (%) | Detection method | Refs |
|---|---|---|---|---|---|---|
| 200 | Modelled ultra-pure water | PFOS and PFOA | C2–C7 | — | LC-MS/MS | [201] |
| 354 | Modelled ultra-pure water | PFOA and PFOS | CO, CO2,
|
50 | HPLC-MSD-ion Trap | [203] |
| 575 | Modelled ultra-pure water | PFOA and PFOS |
|
— | LC-MS/MS | [204] |
| 700 | Groundwater | 6:2 FTAB and HFPO-DA | F‒ | 98 | LC-MS/MS | [205] |
| 700 | Groundwater | 15 PFAS |
|
95–100 | LC-MS/MS | [206] |
| 44, 400, 500, and 1,000 | Modelled ultra-pure water | PFOS | F‒ | 3.4–103.3 | LC-MS/MS | [207] |
| 354 and 612 | Groundwater | PFOA and PFOS | H2 and CO | 39–40 | LC-MS/MS | [208] |
5.3 Reductive chemical processes
5.3.1 Zerovalent metals (ZVM)
The reductive chemical process by ZVM involves the transportation of pollutants to the ZVM surface through mass transfer, where they are adsorbed and transformed into less toxic or nontoxic species, followed by the desorption and mass transfer of by-products into solution [209]. The metal (M) can be aluminium (Al), zinc (Zn), iron (Fe/I), or copper (Cu). The surface properties of the ZVM affect contaminant reactivity since the process is a surface-mediated electron transfer process [210]. Temperature, pH, and ZVI concentration are other factors that influence the effective degradation of PFOS [211]. Studies have been reported based on the ZVM-based reductive decomposition of PFASs and their performance (Table 11) [8,211–215]. The reported studies mostly used iron (ZVI); the reported efficiencies reached 98%. For example, Hori et al. successfully degraded PFOS using zerovalent iron in subcritical water (350°C) [212]. No short-chain PFAS by-products were formed. Only about 0.7% of CHF3 was detected in the gas phase after degradation and F− ions. Compared to other metals, such as Al, Cu, and Zn, using iron led to the most efficient PFOS decomposition [212]. Another investigation by Hori et al. highlighted that the increase in the specific surface area was an important factor in accelerating the decomposition of PFHS [213]. In a study by Blotevogel et al., zerovalent iron (ZVI) and zinc (ZVZn) were used for the reductive oxidation of PFOA [215]. The group reported increased PFOA degradation efficiencies from 50% to 100% over 61 days using ZVI and ZVZn [215]. No intermediates were observed during this period, suggesting that PFOA removal was not because of reductive defluorination by ZVMs. Nevertheless, the terminal carboxylate was reduced by ZVMs [215]. PFOA and PFOS were the most investigated PFAS using ZVM-based reductive degradation (Table 11). The only studied water matrix was modelled ultra-pure water. The most common by-product was F−. Only one study obtained short-chained PFAS as by-products/intermediates [214]. The observation could indicate that the reductive oxidation of PFAS did not occur stepwise, as also observed during sonolytic degradation. Overall, most studies that used ZVI-based degradation of PFAS afforded almost 100% efficiency. The commonly used detection techniques after ZVM-reductive oxidation of PFAS include LC-QTOF-MS, HPLC-CDD, LC-MS/MS, and HPLC-ESI-TOF-MS. Amongst them, HPLC-CDD and LC-MS/MS were the most used.
Analytical performance of ZVM-based reductive degradation methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology | Sample matrix | Target analyte | By-products/intermediates | Degradation efficiency (%) | Detection method | Refs |
|---|---|---|---|---|---|---|
| Mg-aminoclay (MgAC) coated ZVI | Modelled ultra-pure water | PFOA, PFNA, PFOS, and PFDA | F‒ | 38–96 | LC-QTOF-MS | [211] |
| ZVI in supercritical water | Modelled ultra-pure water | PFOS | F‒, CHF3 | >98 | HPLC-CDD | [212] |
| ZVI in Sub- and supercritical water | Modelled ultra-pure water | PFHS | F‒,
|
— | HPLC-CDD | [213] |
| ZVI and ZVI + biochar | Modelled ultra-pure water | PFOA, PFHpA, PFHxA, PFOS, PFHpS, PFHxS, and PFBS | F‒ | 17–94 | LC-MS/MS | [8] |
| ZVI-NPs with or without 1% polyvinylpyrrolidone (PVP) coating | Modelled ultra-pure water | PFOA | PFHpA, PFHxA, PFPeA, PFBA, and F‒ | — | LC-MS/MS | [214] |
| ZVI and ZVZn | Modelled ultra-pure water | PFOA | — | 50–100 | HPLC-ESI-TOF-MS | [215] |
5.3.2 Advanced reduction processes (ARPs)
ARPs combine activation methods such as electron beams, microwaves, ultrasound, and ultraviolet beams with reducing agents such as iodide, sulfite, dithionite, and ferrous iron to form reactive reducing radicals that degrade pollutants to less toxic products. The oxidizing hydroxyl radical (OH˙), the reducing hydrogen radical (H˙), and the hydrated electron (
Analytical performance of advanced oxidative methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology | Sample matrix | Target analyte | By-products/intermediates | Degradation efficiency (%) | Detection method | Refs |
|---|---|---|---|---|---|---|
| UV-KI | — | PFOA, PFOS | Perfluoroalkyl iodides (CxF2x + 1I) | — | LC-MS/MS | [218] |
| UV-KI | Modelled ultra-pure water | PFOA, PFOS | C1–C8 non-iodinated and iodinated gaseous intermediates | — | HPLC-MS | [219] |
| UV-KI | — | PFOA | C2–C7, iodinated and fluorinated hydrocarbons | — | LC-MS/MS | [199] |
| UV-KI | Modelled ultra-pure water | PFOA | C2–C7, iodinated and fluorinated hydrocarbons | 98 | LC-MS/MS | [216] |
| UV-KI | Modelled ultra-pure water | PFOA | C2–C7 | 80.91 | LC-MS/MS | [220] |
| UV-
|
Modelled deionized water | PFOA | C2–C7, fluorinated alkyl sulfonates | 88.5 | HPLC-TQD/MS | [221] |
| UV-
|
Modelled ultra-pure water | PFOA | F− | — | — | [222] |
| UV-β-Ga2O3 | Modelled ultra-pure water | PFOA | F− and shorter-chain PFCAs | 98.8 | HPLC-MS and LC-MS/MS | [217] |
In addition, PFOA was the most studied analyte using this technology compared to other PFAS. Based on the studies reported in Table 12, modelled ultra-pure water is the only analysed water matrix for PFAS using ARP technology, which prompts more research to be conducted on other water matrices, as the conclusions on ARP technology, especially matrix effects, may be biased since different water matrices have not been investigated. The most common by-products or intermediates include C2–C7, iodinated and fluorinated hydrocarbons. Using ARP for PFAS removal, the efficiency was more than 80%, and LC-MS/MS is the most used detection method in analysing PFAS after removal using ARP technology.
5.4 Plasma-based technology
Various advanced oxidation processes (AOPs) based on electrical discharge plasma have been used for water treatment [223]. These include different types of reactors, discharge phases, electrode materials, electrode configurations, and additives. Simultaneous oxidization and reduction of organic molecules can be achieved using plasma electrical discharges [223,224]. While the oxidation method is centred primarily on hydroxyl radical attack, the reduction process may involve synchronized reactions of hot plasma electrons, background gas ions, and aqueous electrons [224]. Several studies have investigated plasma to treat PFAS from water (Table 13) [224–228]. For example, research by Stratton et al. indicated that only 10% of PFOA and PFOS were degraded into shorter chain PFAAs during an electrical discharge plasma procedure. In addition, they reported that typical primary plasma-derived reactive species (hydroxyl and superoxide radicals) were not involved. Instead, the aqueous electrons, free electrons, and argon ions were responsible for degrading the PFAS [224].
Analytical performance of plasma-based advanced oxidative methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology | Sample matrix | Target analyte | Reactive species | By-products/intermediates | Degradation efficiency (%) | Detection method | Refs |
|---|---|---|---|---|---|---|---|
| High-voltage pulsed discharges | Groundwater | PFOS and PFOA | Aqueous electrons, free electrons, and argon | C5–C7 | 10 | UPLC-QToF-MS | [224] |
| High-voltage plasma pulses | Modelled ultra-pure water | PFOA and PFOS | Plasma electrons, aqueous electrons, and argon ions | C4–C7, F‒, and
|
90 | UPLC-QToF-HRMS | [225] |
| DC plasma within oxygen bubbles | Treatment water | PFOA and PFOS | Plasma electrons | C1–C6 and F‒ | 100 | HPLC/MS | [226] |
| DC plasma in air, N2, and Ar microbubbles | Modelled ultra-pure water | PFOA | Aqueous electrons, free electrons and | C3–C7 | 81.5–95.3 | UHPLC-QTOF-MS | [227] |
| DC plasma in Ar, O2, He, and Ne bubbles | — | PFOA and PFOS | ˙OH radicals, e‒ (aq) |
|
— | HPLC-MS | [228] |
Contrastingly, a similar study by Singh et al. reported a 90% degradation of PFOA and PFOS using plasma treatment. But, like Stratton et al., the group said plasma electrons, aqueous electrons, and argon ions were accountable for PFOA and PFOS mineralization. Degradation of PFOS was suggested to have occurred stepwise, with PFOA being the first by-product and PFBA the last. The group indicated that the mechanism of degrading PFOA began at the COOH group, in which PFOA was degraded to form ˙C7F15 radicals. These unstable radicals reacted with ˙OH radicals to generate an unstable alcohol (C7F15OH) that was subsequently mineralized to produce a stable ketone (C6F13COF) by removing HF via the thermal transfer of aqueous electrons. The breakdown of PFOS seemed comparable to PFOA’s, except the study reported the C–S bond being cleaved from the terminal carbon by electrons or argon ions during the initiation step to form SO3
− group and ˙C8F17 radicals. Subsequent reductive, oxidative, and hydrolysis chain propagation reactions generated short-chain PFCAs as outlined in the mechanism (Figure 5) proposed by the group below [225]. Unfortunately, most studies on AOP technology for PFAS removal are limited to PFOA and PFOS only (Table 13). Although the scope of the investigation is defined in terms of the variety of PFAS analysed, at least the water matrix scope investigated is wide, including groundwater, modelled ultra-pure water, and treatment water [224–227]. Aqueous and free electrons were the most common reactive species responsible for mineralizing PFOA and PFOS in most reported studies. The degradation efficiency was above 80% for most studies, except for one study by Stratton et al. as noted prior, where a very low removal of 10% was achieved. Most studies reported only fluorinated carbon chains ranging between C3 and C7 as by-products, while others included F− or
![Figure 5
Reported mechanism of degrading PFOA and PFOS using plasma technology and their by-products. Reproduced with permission from the study by Singh et al. [225].](/document/doi/10.1515/revac-2023-0064/asset/graphic/j_revac-2023-0064_fig_005.jpg)
Reported mechanism of degrading PFOA and PFOS using plasma technology and their by-products. Reproduced with permission from the study by Singh et al. [225].
5.5 Adsorption
Adsorption can be defined as the attachment of molecules to a solid material’s surface through physical or chemical interactions. Examples of solid materials used in PFAS investigations include AC, covalent triazine-based framework (CTF), metal-organic frameworks (MOFs), carbon nanotubes, ash, char, polyaniline, molecularly imprinted polymers (MIPS), and metal oxides (Table 14) [229–238]. Amongst these, AC is the most used adsorbent for removing PFAS. It is produced from organic materials with high carbon content, such as wood, coal, and lignite [236,237]. Of the two forms of AC, granular activated carbon (GAC) has been reported to remove almost 100% PFAS from drinking water compared to powder activated carbon (PAC). More efficiently, longer-chain PFAS (e.g., PFOS and PFOA) than shorter-chain PFAS (e.g., PFBS and PFBA) [236]. Despite its advantages, AC incurs problems of subsequent difficult and expensive costs to clean up the produced residue after use and regeneration costs [231,236]. Coagulation, flocculation, and sedimentation are usually required to recover the spent PAC from the solution resulting in complex compositions of the sediments [231]. To solve this problem, magnetic nanoparticles such as iron oxide (Fe3O4) are incorporated into the GAC or PAC surface [231]. According to Du et al., pH and competitive pollutants in actual wastewater samples influence the removal efficiency of the AC adsorbent [229]. In an investigation by Wang et al. using a CTF comparatively with pulverized microporous AC and single-walled carbon nanotubes to remove PFCAs from aqueous solution [232]. CTF was reported to have higher adsorption capacity and binding affinity among the used adsorbents. Strong electrostatic interactions between the protonated triazine groups and the negatively charged PFAAs head groups were the favoured mechanism of adsorption [232]. A few studies have reported using adsorption for removing PFAS in water systems [116,229–235,238]. Table 14 shows that modelled ultra-pure water, groundwater, and wastewater are the most analysed water matrices using adsorption. A variety of adsorbent materials, which include F-COPs, bamboo AC, CMF50%@PANI, Fe3O4-AC (MAC), CTF, MOF, PANF-derived PACFs, maize ash, SWCNT, and char, have been used. Amongst the analysed PFAS, namely PFOA, PFNA, PFOS, PFDA, PFUnDa, PFDoDa, PFTeDa, PFHxA, PFHpA, PFBS, and PFHxS, most studies focused on PFOA or PFOS. The most prevalent detection methods for analysis were LC-MS/MS and HPLC-CDD after removal. For most of the studies, the kinetic model followed a pseudo-second-order rate. Also, most studies did not report the number of reusability cycles except for a few that could reuse their adsorbent materials more than three times [229,231,232,234].
Analytical performance of adsorption-based methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Adsorbent | Sample matrix | Target analyte | Adsorption efficiency | Detection method | Kinetic model | Isotherm model | Sorption–desorption cycles | Refs |
|---|---|---|---|---|---|---|---|---|
| F-COPs | Modelled ultra-pure water | PFOA, PFNA, PFOS, PFDA, PFUnDa, PFDoDa, and PFTeDa | 92.3–100 mg·g−1 | UPLC-MS | Pseudo-second order | Langmuir | — | [116] |
| Bamboo AC | Wastewater | PFHxA, PFHpA, PFOA | 0.06–2.82 mmol·g−1 | LC-MS/MS and HPLC-CDD | Pseudo-second order | Langmuir | 5 | [229] |
| CMF50%@PANI | Modelled ultra-pure water | PFOA and PFOS | 46.71 and 47.36 | XPS analysis | Pseudo-second order | Langmuir | — | [230] |
| Fe3O4-AC (MAC) | Modelled ultra-pure water | PFOS, PFBS, PFOA, PFHxS | 0.021–1.63 mmol·g−1 | HPLC-CDD | Pseudo-first-order | Freundlich and Langmuir | 6 | [231] |
| CTF | Modelled ultra-pure water | PFOS and PFOA | — | HPLC-CDD | — | Langmuir | 4 | [232] |
| MOF | Groundwater | 28 PFAS | — | UPLC-MS/MS | — | — | — | [233] |
| PANF-derived PACFs | Water | PFOS and PFOA | 1.52 and 0.73 mmol·g−1 | HPLC-CDD | Pseudo-second order | Freundlich | 5 | [234] |
| Maize ash | Groundwater and wastewater | PFOS | >700 mg·g−1 | LC-MS/MS | Pseudo-second-order model | Langmuir | — | [235] |
| SWCNT | Groundwater and wastewater | PFOS | >700 mg·g−1 | LC-MS/MS | Pseudo-second-order model | Freundlich | — | [235] |
| Char | Groundwater and wastewater | PFOS | <170 mg·g−1 | LC-MS/MS | Pseudo-second-order model | Freundlich and Langmuir | — | [235] |
5.6 Anion exchange
Positively charged anion exchange resins (AERs) attract and hold negatively charged pollutants such as PFAS to successfully eradicate them from water systems [239]. Although AER has been reported to have a high removal capacity for many PFAS, it is costly compared to adsorption by GAC [239]. In water matrices with high levels of inorganic anions and hydrophobic non-organic matter (NOM), adsorption site competition and pore blockage exist, which lowers PFAS adsorption [240,241]. In addition, microbial growth and inorganic particulate can block pores, leading to high column head loss and decreased usable bed life [241]. AERs are typically reactivated using salt (NaCl, KCl, and NH4Cl) and alkaline (NaOH, KOH, and NH4OH) aqueous solution to desorb inorganic anions and NOM via rinsing and backwash [241,242]. Although backwash works for GAC, it is not advised for AER as it may result in the poor arrangement of uniform resin beads, which could disturb bed mass transfer zones due to a disrupted bed [240]. Chow et al. observed this effect in AER breakthrough curves when bed disruptions happened (Figure 6) [241]. Also, the group reported that oxidants such as chlorine might damage the resin and produce nitrosamines, especially for the gel AERs used in their investigation. In a similar study by Zaggia et al., backwashing of AERs using dilute solutions of 0.5% NH4OH and 0.5 NH4Cl was reported to have enhanced the removal capacity of a hydrophobic A532E resin [242]. The group reported higher removal capacity for A532E compared to the non-hydrophobic A600E and mildly hydrophobic A520E [242]. Despite its efficiency in removing PFAS, the major drawback of AERs, as reported by Chow et al., is its limitation regarding the broad-range removal of PFAS, where different co-pollutants exist, since AERs are highly selective for anionic PFAS [241].
![Figure 6
Influent-normalized breakthrough of linear PFAS in each tested column at effluent port (GAC EBCT = 10.3 min; AER EBCT = 2.7 min). Modelled dose–response breakthrough curves are shown with measured data for visual aid. Breakthrough at 10%, 50%, and 100% are shown for reference. Dotted red vertical lines correspond to BVs of bed disruptions from pilot interruptions that led to discontinuous breaks in some breakthrough curves at days 111 and 264, reproduced with permission from the study by Chow et al. [241].](/document/doi/10.1515/revac-2023-0064/asset/graphic/j_revac-2023-0064_fig_006.jpg)
Influent-normalized breakthrough of linear PFAS in each tested column at effluent port (GAC EBCT = 10.3 min; AER EBCT = 2.7 min). Modelled dose–response breakthrough curves are shown with measured data for visual aid. Breakthrough at 10%, 50%, and 100% are shown for reference. Dotted red vertical lines correspond to BVs of bed disruptions from pilot interruptions that led to discontinuous breaks in some breakthrough curves at days 111 and 264, reproduced with permission from the study by Chow et al. [241].
In a study by Gao et al. who investigated the removal of PFOS and potassium 2-(6-chloro-1,1,2,2,3,3,4,4,5,5,6,6-dodecafluorohexyloxy) (F-53B) from wastewater using IRA67 resin, the removal capacities were 5.5 and 4.2 mmol·g−1, respectively [243]. The group reported that the adsorption mechanisms involved the formation of micelles, hydrophobic interactions, and ion exchange [243]. Similar studies are reported in Table 15. Many studies targeted PFOA and PFOS, while others included PFBS and PFBA. The most analysed water matrices were groundwater and drinking water using AERs. After PFAS removal from water using AER technology, the detection methods were HPLC-MS/MS, HPLC-CDD, TOP essay, HPLC-HRMS, and HPLC-ESI/MS/MS. Amongst them, HPLC-MS/MS and HPLC-CDD were the most common. A few studies that regenerated the resin used NH4Cl/NH4OH, NH4Cl/NaCl, NaOH/NaCl, and CH3OH mixtures.
Analytical performance of anion exchange-based methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology (resin) | Sample matrix | Target analyte | Detection method | Regeneration solution | Refs |
|---|---|---|---|---|---|
| Amberlite IRA-400 | Model ultra-pure water | PFOS and PFOA | LC-MS/MS and HPLC-CDD | — | [232] |
| A520E and A860 (Purolite) | Groundwater | PFOA and PFOS | LC-MS/MS | — | [240] |
| PFA694E, PSR2 Plus | Groundwater | PFCA | LC-MS/MS, TOP essay, and HPLC-HRMS | — | [241] |
| A600E, A520E and A532E | Drinking water | PFOA, PFOS, PFBA and PFBS | — | NH4Cl/NH4OH | [242] |
| A600E, PAD500, A520E, and MN102 | Groundwater and drinking water | PFBS, PFBA, PFOA and PFOS | HPLC-ESI/MS/MS | NH4Cl/NaCl | [244] |
| A-600 (Purolite) | Drinking water | 14 PFAS | LC-MS/MS | — | [245] |
| IRA910 | Groundwater | PFHxS | HPLC-CDD | — | [246] |
| IRA67 | Wastewater | F-53B and PFOS | HPLC-CDD | NaOH/NaCl and CH3OH | [243] |
5.7 High-pressure membranes
High-pressure membranes, including nanofiltration (NF) and reverse osmosis (RO), have been reported as effective for treating PFAS in water [247]. RO rejects salts to a high degree, while NF membrane rejects hardness to a high degree. RO has tight membranes, and its technology depends on membrane permeability. Generally, NF membranes have higher rejection potential than microfiltration and ultrafiltration membranes [248]. They also have a higher water flux than RO [248]. Unlike adsorption using anion exchange or GAC, removal efficiencies of membrane processes do not depend on the concentrations of salts, organic matter, or co-pollutants. Their mechanism of rejection is based on size exclusion [249]. Normally, the quality of water samples differs for different water matrices, which affects the retention of small trace organic compounds (TOrCs). For example, the effect of solute concentration in rejecting TOrCs varies depending on the nature of the TOrCs and NF membrane. Increasing the ionic strength may cause electrostatic shielding of several charges from the NF membrane and charged TOrCs, reducing the retention of TOrCs.
In addition, high ionic strengths are likely to shrink the membrane pores, increasing the rejection of contaminants due to the improved size exclusion [250]. However, extreme ionic strengths may cause membrane pores to swell, causing TOrCs rejection to decrease [250]. In addition to concentration and ionic environment, properties such as hydrophobicity, charge, and molecular size affect TOrC rejection by NF membranes [250]. Membranes can be operated for long periods if membrane fouling is controlled in the treatment train design [249]. When water samples contain very large organic matter, membrane fouling is unavoidable and occurs stepwise. For smaller TOrCs, the NF membrane surface properties continuously change until a lasting fouling layer of attached organic matter is formed [250]. Many studies have investigated the treatment of PFAS from water systems using NF and RO [248–254]. As shown in Table 16, a wide range of water matrices have been analysed, including contaminated effluents, groundwater, synthetic groundwater, modelled ultra-pure water, and wastewater. PFAS analysed using membrane technology included PFOS, PFBA, PFAS in AFFF, perfluoroalkyl ether acid (PFEA), TFS, and PFPeA. Amongst them, PFOS was the most targeted based on Table 16. The average efficiency of membrane-based technologies is impressive since most studies reported more than 85% removal of PFASs [248–251,253,254]. However, membrane fouling was reported in some studies [249,253]. The detection methods used after removal included LC-MS/MS, HPLC-QToF-MS, and UPLC-MS/MS. A study by Wang et al. using NF to remove PFOS and PFBS from water reported 88% retention of PFOS in all water samples, which was higher than the 69% retention obtained for PFBS [250]. These results could be attributed to the stronger hydrophobicity and larger molecular size of PFOS. Moreover, membrane pores were blocked at higher solute concentrations, which enhanced size exclusion and consequently increased the rejection of PFOS and PFBS [250]. The group reported that the increasing ionic strength from 0 to 100 mM improved PFOS retention from 89.6% to 91.9%. On the contrary, PFBS rejection was reduced from 48.9% to 20.5%. From the aforementioned results and with a decrease in membrane charge and flux when ionic strength was increased, it was suggested that electrostatic repulsion and size exclusion were the primary mechanisms governing the rejection of both analytes [250]. Liu et al. reported that the rejection of 10 PFAAs and 32 PFASs identified in AFFF from groundwater using NF was between 92% and 98% and was influenced by the background water matrix constituents [251]. A similar study by Zhi et al. reported that rejections of fluorotelomer sulfonates (FTSs) and PFEAs by DK membranes increased from 88.3% to 97.1% and 81.7% to >99.9% using DK membranes [252]. The study reported that the rejection of PFAS was influenced by steric hindrance and hydrophobic interactions [252]. In addition, Chen et al. used direct contact membrane distillation to concentrate and remove PFPeA from water [253]. The group reported severe membrane fouling and that increasing temperature decreased the rejection of PFPeA from 85% to 58% [253].
Analytical performance of high-pressure membrane methods coupled with various analytical detection techniques for the analysis of PFAS in environmental matrices
| Technology | Sample matrix | Target analyte | Rejection efficiency (%) | Detection method | Membrane fouling | Refs |
|---|---|---|---|---|---|---|
| Nanofiltration | Contaminated effluents | PFOS | 93–97.5 | — | — | [248] |
| Nanofiltration | Water | PFOS and PFBA | 88, 69 | LC-MS/MS | Yes | [250] |
| Nanofiltration, reverse osmosis | Groundwater | PFAS in AFFF | 92–98 | HPLC-QToF-MS | — | [251] |
| Nanofiltration | Synthetic groundwater | PFAS, PFEA, and TFS | 66–99.9 | UPLC-MS/MS | — | [252] |
| Direct contact membrane distillation | Modelled ultra-pure water | PFPeA | 85 | — | Yes | [253] |
| Reverse osmosis | Wastewater | PFOS | 99 | LC-MS/MS | — | [254] |
6 Conclusions and recommendations
The recent increase in the detection of PFAS in water systems worldwide has resulted in the development of sensitive sample preparation techniques that can detect and improve analyte analysis even at trace concentrations. The number of publications reporting on SPE-based technology for separating PFAS in water suggests that the method is most preferred for sample preparation. This review summarized the recent application of various sample preparation and removal methods for selective extraction, preconcentration, and removal of different PFAS in water systems. Multiple studies showed that using SPE-based technologies, especially MSPE, allows for improved sensitivity using modified adsorbent materials and easy, fast recovery of the adsorbent after application. From the reported removal techniques, photochemical oxidation was the only technique that could consistently produce 100% removal of PFAS as observed from multiple studies. Although adsorption is generally preferred for its easy design, cheap operation, and sorbent costs, oxidative and reductive processes are also good in that some are able to degrade PFAS completely, while others can at most breakdown longer PFAS into their shorter counterparts. Their major downside as observed in various tables is the resulting by-products, especially toxic by-products and ones that subsequently cause secondary pollution. Furthermore, the literature shows that carbon-based adsorbent materials are used most for extracting and removing PFAS in water. Amongst other PFAS, the literature indicates that PFOA and PFOS are still the most investigated PFAS for preconcentration and removal studies. Although the literature shows that PFAS can be detected and quantified even at very low concentrations in water, the detection efficiency is reduced by competition or interfering species. Recent studies have developed MIPS to allow for specificity and minimize competition with other contaminants, for example, in the analysis of PFAS in milk [255]. MIPS could be used for PFAS analysis, especially in real water samples. The reports on the effective use of MIPS for extracting PFAS in other mediums suggest that MIPS will likely form a part of the next generation of adsorbent materials for SPE procedures. Moreover, the development of magnetic MIPS for use as adsorbent materials during SPE for extraction and preconcentration of PFAS and for exploration in removing PFAS in water offers science an advanced method with benefits such as easy recovery, fast removal, improved selectivity, increased sensitivity, and low cost. Modifications are still needed to design methods that can eliminate all types of PFAS in water, owing to the literature indicating that some PFAS, especially PFOS for some techniques, could not be degraded. In such cases, improvements are required.
Acknowledgments
The authors wish to acknowledge the financial support from the Department of Science and Innovation-National Research Foundation South African Research Chair Initiative (DSI-NRF SARChI) funding instrument, and Sasol postgraduate bursary. Also, the authors would like to thank the University of Johannesburg, Faculty of Science, and Department of Chemical Sciences for laboratory space.
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Funding information: This study was funded by National Research Fund (NRF) South African Research Chairs Initiative (SARChI) funding instrument (grant no. 91230) and Sasol.
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Author contributions: Nompumelelo Malatji: conceptualisation of the work; formal analysis investigation, writing – original draft. Anele Mpupa: conceptualisation of the work, supervision, validation, writing – review and editing. Philiswa Nosizo Nomngongo: conceptualisation of the work, funding acquisition, supervision, validation, and writing – review and editing. All listed authors have revised and approved the submission of the manuscript.
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Conflict of interest: Authors state no conflict of interest.
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Data availability statement: The datasets generated during and/or analyzed during the current study are available from the corresponding author on reasonable request.
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- Special Issue: Greener Analytical Methods - Part I
- Green novel photometric and planar chromatographic assays of remdesivir: Comparative greenness assessment study using estimated GAPI tool versus ISO technical reported methods
- Enhancing methane sensing with NDIR technology: Current trends and future prospects
- Eco-friendly HPLC method by using response surface design to measure a combination of three antidiabetic drugs
- Metal and metalloid monitoring in water by passive sampling – A review
- Green chemistry and its implementation in pharmaceutical analysis
- Synthesis of a benzothiazole-based structure as a selective colorimetric-fluorogenic cyanate chemosensor
