A38. 20240819 brief OP aan gemeenteraden PFAS schade.pdf

Thus, more research is needed to explore the stability of PFAS in S/S-soil over time. 6.3 Destruction of PFAS in Soil Using Thermal Treatment Treatment of contaminated soil by exposure to high temperatures (850-1200 °C) may be an effective technique to mineralize PFAS [96]. For example, incineration in a fluidized bed incinerator at temperatures of up to 900 °C can reduce the PFOS levels to below the detection limits in the exhaust air (15 ng/m3), while also reducing solid residues (10 µg/kg dry matter) and water discharges (25 ng/L) [96]. However, incineration of PFAS-contaminated soil together with other wastes may lead to the formation of highly volatile and mobile 1H-perfluoroheptane, dioxins, furans, tetrafluoromethane, and hexafluoroethane (C2F6) [97-99]. Thus, the addition of appropriate substances, such as Ca(OH)2, may be required to reduce the emission of these harmful by-products [100]. More research is needed to better understand the effects of incineration on PFAS and the byproducts formed. Thermal treatment is not suitable for large masses of soil contaminated with PFAS to be subsequently landfilled due to the high costs involved [9]. 7. Treatment of Landfill Leachate Contaminated with PFAS In modern landfills, leachate is collected and treated on-site at local treatment plants and then discharged into a recipient water body or pumped to WWTPs [65].

Particularly in Sweden, the landfill leachate must be treated before it is delivered to WWTPs to meet the quality requirements set by the WWTP operators. Most of the traditional treatment systems for landfill leachates are not designed to remove PFAS, and thus, the extent to which they reduce the PFAS levels is not wellknown due to a lack of published data. The processes for removing PFAS from landfill leachate discussed in the literature include separation of PFAS from leachate by sorption, biological treatment, membrane-based treatments (such as nanofiltration and reverse osmosis), and compacted wetlands (Table 3). Destruction techniques used for removing PFAS from groundwater, such as electrochemical treatment, thermal treatment, wet air oxidation, and advanced oxidation processes (such as photolysis, photocatalysis, and activated persulfate oxidation), have also been discussed.

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7.1 Separation of PFAS from Landfill Leachate 7.1.1 Sorption The sorption of PFAS to GAC (e.g., Filtrasorb 300 and 600, or AquaCarb 1240) and anion exchange (AE) materials (such as Purolite® A600) has been tested for the treatment of landfill leachate, groundwater, and drinking water [1, 58, 101]. These sorbents are more effective for eliminating long-chain PFAS than short-chain PFAS (e.g., PFOS>PFBS, in terms of removal efficiency), but they have a low sorption capacity [47, 76, 85]. As shown in Table 3, Busch et al. [1] and Yu et al. [76] found that AC can adsorb 68.2% to 99.8% of PFAS. However, in these studies, ∑PFAS in the treated leachates remained high (e.g., 4,079 ng/L) due to the high concentrations in raw leachate (e.g., 12,819 ng/L) and the short residence times of the leachate through the AC filters, or high mass flows. Additionally, DOC in leachates may compete with PFAS for the adsorption sites. Short-chain PFAS, such as PFBA (C4), PFPeA (C5), and PFHxA (C6), maybe inefficiently adsorbed because they are less hydrophobic compared to long-chain PFAS, such as PFNA (C9), PFDA (C10), and PFOS (C8), and may, therefore, be displaced by long-chain PFAS and DOC [58].

Thus, the selective nature of PFAS removal and the associated desorption of short-chain PFCAs during the co-removal of multiple PFAS must be considered during the designing and operation of the adsorption treatment processes. 7.1.2 Membrane Technology Reverse osmosis (RO) and nanofiltration (NF) involve the use of semi-permeable membranes (with 1-10 nm and <1 nm pores, respectively) to remove contaminants from leachates. These are physical separation technologies that transfer PFAS from the leachate to a secondary phase (rejected water or concentrate), which then needs further treatment. Appleman et al. [101] reported that NF removed PFOS, PFOA, PFDA, PFPeA, and PFBA to below-detection limits (with >97% removal efficiency) from artificial groundwater spiked with 1 µg/L of each PFAS; PFBS and PFHxS were present in extremely low concentrations (e.g., 10 and 20 ng/L, respectively) in the permeate. Soriano et al. found that the NF90 membranes with a smaller pore size (0.34 ± 0.02 nm) rejected over 99% of PFHxA at 100 mg/L from the industrial effluents [102]. Their results showed that size exclusion and electrostatic interaction occurred simultaneously and affected the selectivity of PFHxA separation.

Reverse osmosis in WWTP was able to remove PFHxA, PFHpA, PFOA, PFNA, PFDA, PFBS, PFHxS, and PFOS below the detection levels from their initial concentrations, which varied between 3 µg/L for PFNA and 39 µg/L for PFOS [103]. After membrane-treatment, PFAS were concentrated in the rejected water, which was about 15% of the treated water volume for RO [103]. Soriano et al. showed an efficient removal of PFHxA (over 99%) using XLE and BW30, two types of RO membranes [102]. These results indicate that both NF and RO can be effective in treating PFAS of a wide range of molecular weights. Franke et al. removed PFAS from drinking water using a combination of NF with AE or GAC adsorbents [104]. In the cited studies, water with no DOC was used, while the landfill leachates may contain significant amounts of DOC. Thus, more research is required on the membrane filtration treatments of PFAS in landfill leachate. Although RO and NF showed promising results in removing PFAS from contaminated water, Fuertes et al. [71] found a two-fold increase in the concentration of PFAS in the membrane bioreactors of the effluent treatment plants (using ultrafiltration treatment) compared to the raw leachate, indicating

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degradation of the precursor compounds, which increased the levels of some PFAS [71]. These results indicated that not all membrane methods are suitable for treating PFAS-contaminated water. 7.1.3 Biological Treatments Biological treatments (BIO) applied by Busch et al. [1], Yan et al. [3], and Fuertes et al. [71] were found to be ineffective in removing PFAS from landfill leachate. Instead, the concentrations of PFCAs and PFSAs were increased with the use of these types of treatments (Table 3). This could have occurred due to the resistance of PFAS to biodegradation owing to their high-energy carbon-fluorine bonds [105], and the possible biodegradation of precursor compounds (such as N-ethyl perfluorooctane sulphonamidoethanol and fluorotelomer-based compounds, or other unidentified components) to PFCAs and PFSAs, as observed for the biological processes in WWTPs [7, 19, 106]. Constructed wetlands (CWs) include various kinds of aerated lagoons, sedimentation tanks, reed beds, and polishing ponds. Multiple remedial processes may be conducted simultaneously in these systems, including biodegradation (aerobic and anaerobic), phytoremediation (plant uptake and enhancement of biodegradation), and sorption on diverse substances [87]. During tests in a large tropical wetland, 61% of ∑PFAS and 50%-96% of individual PFAS were removed from landfill leachate.

The most effective processes were sorption of long-chain PFAS to soils and sediments and uptake of short-chain PFAS by plants. These processes removed about 44% of ∑PFAS during the passage of the leachate through a reed bed. Additionally, an aeration lagoon removed 55%-73% of the PFAS precursors through biotransformation, but it was ineffective in the removal of long-chain PFAS. This approach may only be suitable as a pre-treatment for landfill leachate as the concentrations of ∑PFAS in the treated leachate were still high (Table 3). However, the use of young plants or species with a high affinity for PFAS, and the substrates with high PFAS sorption capacity can potentially improve the removal of PFAS from landfill leachates in CWs. 7.2 Techniques for the Destruction of PFAS in Landfill Leachate 7.2.1 Oxidation Processes Chemical oxidation involves the use of oxidants, such as sodium permanganate (NaMnO4), sodium persulfate (Na2S2O8), hydrogen peroxide (H2O2), or ozone (O3) to degrade organic pollutants and decontaminate water. For example, H2O2 mixed with water generates hydroxyl radical (OH•), which is an extremely strong oxidant (E = 2.33 V) that reacts with organic molecules with the reaction rate-constants ranging between 106 and 1010 Ms–1 [107]. H2O2 is often used in conjunction with Fe2+ as a catalyst.

The efficiency with which the chemical oxidation processes degrade PFAS strongly depends on the pH, temperature, and initial concentration of the oxidant. Ozone does not degrade PFOA and PFOS significantly at pH 4-5 [108], but NaMnO4 has been shown to degrade 47% of PFOS in an aqueous solution at pH 4.2 in 18 days. Increasing the pH to about 10-11 (by adding NaOH to the solution) decreases PFOS degradation to about 10% of the that achieved by treatment with NaMnO 4, while the addition of 8.7 g O3/h (2.5 wt%) reduces the concentrations of PFOA and PFOS from 50 µg/L to about 5 µg/L and 7.5 µg/L, respectively, after 4 h [108, 109]. Ozonation is less efficient without pre-treatment; the results are best when O3 is added at an acidic pH and the process is continued by adding O3 at an alkaline pH [108]. At an acidic pH, O3 Page 16/30

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decomposes slowly via reaction with hydroxide ions (as their concentrations are low), and dissolved O3 accumulates in the solution. When the concentration of O3 is sufficiently high in the solution (e.g., 3.5 mg/L), making the solution alkaline, it generates large amounts of OH• that react rapidly with O3 and form abundant reactive free radicals, including potent hydroxyl radicals that can efficiently degrade PFOA and PFAS. Accumulation of O3 at an alkaline pH (without pre-treatment) in the solution is poor (e.g., about 0.3 mg/L at pH = 11) as the large amounts of OH• continually react with O3, causing the concentrations of reactive free radicals to be low throughout the process [108]. The process can be improved by increasing the dose of the oxidant and the temperature (e.g., the kinetic rate constant of PFOS increased from 0.46 × 10–2 L/d at 25 °C to 3.64 × 10–2 L/d at 65 °C) in the tests or when combined with other processes, such as photolysis [92]. A synergistic effect was observed when ozone was combined with UV or with air fractionation, whereby the removal of PFAS reached 73% and 95%, respectively [108]. Heterogeneous photocatalysis is a promising solution for the removal of PFAS from contaminated water.

This technique directly uses photons (hv) from a VUV or UV to generate negatively-charged electrons (e–) and positively-charged holes (h+), which have high oxidation capacity [110]. The positively-charged holes react with water, yielding OH• that can effectively degrade diverse organic compounds. The negatively-charged electrons migrate to the surface of the photocatalyst and react with the adsorbed water to form hydrated electrons (e–aq). PFAS sorbed on the catalyst’s surface are attacked by these hydrated electrons and degraded into shorter-chain compounds [110]. Photocatalysis with TiO2 may degrade PFCAs (e.g., PFOA, PFNA, and PFDA), but several hours are required for satisfactory results. For example, Panchangam et al. reported that TiO 2 irradiated with a 16 W (254 nm) low-pressure mercury lamp could degrade 99% and mineralize 38% of PFOA and PFDA in 7 h [111]. The treatment time could be reduced (several-fold) by adding acid to the treated solution, for example, adding 0.225 M perchloric acid (HClO4) reduced the pH of the solution to 3 and cuts down the treatment time for PFOA from 24h to 7h (Ibid.). According to Panchangam et al., the presence of acid improves the ionization of PFCAs by increasing the chances of electron transfer from PFCAs to the photoholes of excited TiO2, thereby prolonging the life of the photoholes [111].

The process can be further improved by the recovery of the acid. Furthermore, a recent development on the photocatalytic degradation of PFOA using a TiO 2-rGO (0.1 g L–1) composite catalyst was shown to degrade up to 93.7% of PFOA compared to the TiO 2 photocatalysis (24% ± 11%) and direct photolysis (58% ± 9%) [112]. The degradation efficiency was related to several parameters, which included the initial PFAS concentration, photocatalyst, UV lamp intensity, and media. β-Ga2O3 is relatively more efficient as a photocatalyst than TiO2 for PFCA degradation owing to the wider band-gap of the former (4.8 eV for βGa2O3 compared to 3.2 eV for TiO2) and the higher position of its conduction band relative to the vacuum energy level (e.g., ECB values: - 2.95 eV and - 4.21 eV, respectively) [113]. For example, Shao et al. found that PFOA was completely degraded after 65 min of exposure to upgraded β-Ga2O3 along with VUV irradiation at 185 nm and pH 4.3 [114]. The catalytic β-Ga2O3 had a sheaf-like nanostructure consisting of plates elongated in the [1 0 0] direction, a specific surface area of 36.1 m2/g, and high densities of the nanopores (2-4 nm and 8 nm). In contrast, commercial β-Ga2O3 degraded only 38% of the PFOA in 3 h (which was still two times higher than the rate obtained with P25 TiO2) [114].

On the other hand, boron nitride (BN) was found to be quite an efficient catalyst, being 2 to 4 times more active than TiO 2 [115]. Duan et al. Page 17/30

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[115] showed that low concentrations of PFOA were detected after 240 min of photolytic degradation using BN [115]. The removal of PFAS from landfill leachate through photocatalysis may be affected by the presence of dissolved organic matter (DOM), bicarbonate, and chlorides. The DOM and bicarbonate in leachates probably impair the photodegradation process by occupying the catalyst’s surface, thereby reducing its adsorption capacity for PFAS [110]. Thus, applying photocatalysis as the final step may be an effective strategy in leachate treatment (when most of the DOM has been removed in the previous steps). Another option may be to add O3 gas to promote the removal of long-chain PFAS (e.g., PFOA), i.e., by increasing the rate at, which photo-produced H+ reacts with PFOA, rather than recombining with photo-produced e– [112]. Additionally, the pH of the leachate should probably be adjusted to about 4 to promote high adsorption of PFAS and enhance the performance of e– and H+ pairs [110]. Moreover, future research should include the evaluation of the efficiency of photocatalysis in treating short-chain PFAS in landfill leachates.

In electrochemical treatment (ET), “non-active” anodes composed of materials such as Ti/RuO2, SnO2, PbO2, and boron-doped diamond (BDD) are used for degrading and mineralizing nonbiodegradable organic compounds, such as PFOA, PFBA, PFHxA, PFDA, PFBS, PFHxS, and PFOS [116118]. The electrochemical technique is reported to efficiently remove PFAS from wastewater and groundwater, but the required treatment time might be quite long. For example, Gomez-Ruiz et al. reported that ET treatment with a BDD electrode removed 99.7% of diverse PFAS from wastewater [119], with their total concentration dropped from 1,652 µg/L to 4.2 µg/L during 10 h of treatment at a current density of 50 mA/cm2 and a voltage between 13.9 and 15.3 V. Similarly, Zhuo et al. found that a current density of 0.59 mA/cm2 and an oxidation potential greater than 2.76 V were adequate for removing about 97.5% of PFOA from spiked water (at the concentrations of 20, 30, and 50 mg/L) after 2 h of electrolysis [120]. Besides the electrochemical parameters (such as current density and voltage), the success of ET depends on the pH and the initial concentrations of PFAS. The degradation rate of PFOA is higher at an acidic pH than at an alkaline pH. At alkaline pH, OH – migrates to the anode surface due to electrostatic attraction, which limits the sites available for degraded CF3(CF2)6COO– anion [120].

The removal ratio was shown to increase with an increase in the initial PFOA concentration. Thus, ET may be an effective treatment for water with high levels of PFAS contamination at an acidic pH. Regarding the treatment of landfill leachate, Witt et al. recently demonstrated that 98% of PFOA was removed electrochemically under a flow through a BDD cell in a 20 L pilot water treatment system [121]. However, this technology is not fully developed for commercial applications, partly due to high energy demands, which increases the costs of operation. For example, 99.7% removal of PFAS from contaminated water would require approximately 256 kWh/m 3 (corresponding to about 14 euro/m3) [119]. Thus, large-scale implementation of ET would require pre-treatment of the PFAS-contaminated water (e.g., using membrane separation) and the prevention of its mixing with diluting water streams (to reduce the volumes requiring treatment). 7.2.2 Thermal Treatment The thermal treatment methods for water include sonochemistry, sub-critical and supercritical water treatment, microwave-hydrothermal treatment, and wet air oxidation (WAO). Page 18/30

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Sonochemical treatment refers to the use of ultrasonic waves at the frequencies of 20-1000 kHz to form cavitation bubbles in water, and hence, high vapor temperatures (near 4,000-5,000 K) in the vapor core. Thus, PFAS may be pyrolytically decomposed at bubble–water interfaces [122]. Decomposition of PFOA and PFOS results in the production of mineralized fluoride (F –), sulfate (SO42–), CO, and CO2 [123]. Campbell and Hoffmann showed that the treatment’s efficiency relied on the frequency of the ultrasonic radiation and the power density [124]. The measured rate constants of PFOS, PFOA, PFHxA, PFHxS, and PFBS showed that degradation increased linearly with the increase in power density (e.g., the rate constant for PFOS increased from 0.01 L/min at 100 W/L to 0.045 L/min at 330 W/L at the same frequency of 358 kHz). Similarly, the rate of PFAS degradation was shown to increase with an increase in the frequency (e.g., the rate constant for PFOS increased from 0.045 L/min at 358 kHz to 0.06 L/min at 610 kHz at the same power density of 330 W/L). Moreover, the degradation rates of PFOA and PFOS were enhanced by 12% and 23%, respectively, through the application of a dual-frequency (20 and 202 kHz).

The increase in degradation with the application of dual-frequency may be attributed to a better overlap of acoustic waves (which enhances bubble expansion and shortens collapse time), thereby strengthening the cavitation effects and the induction of sonochemical reactions [125]. The treatment cost can be reduced by combining sonification with oxidation, for example, by using persulfate (S2O82–) oxidant to generate sulfate radicals (SO4•–), which increases the efficiency of the process [126-128]. Nevertheless, the higher cost of the sonochemical treatment process compared to other destructive methods may limit its application in the treatment of PFAS [129]. According to Hori et al., pure subcritical water treatment cannot degrade PFOS, but it can degrade PFOS efficiently when used in combination with the addition of zero-valent metals (e.g., Al, Cu, Zn, Fe) to the reaction system [125]. Zero-valent iron, Fe (0), provided the best synergistic results for the degradation of PFOS, followed by Zn, Cu, and Al (in decreasing ability of degradation), in the tests performed by the researchers. For example, adding Fe(0) powder (>99.9%, <53 µm at 4.8-19 mmol concentrations) to subcritical water treatment mixtures reduced the concentrations of PFOS from 93-372 µM to <2.2 µM (at 350 °C, 22 MPa after 6 h), with the formation of F– ions, 46%-51% and no PFCA formation (Ibid.).

Additionally, shorter C-F chains, such as PFHxS, could be degraded from the initial concentrations of 741 µM to 97 µM using subcritical water treatment combined with the addition of 19.3 mmol Fe [130]. This process can be described as the adsorption of PFAS to the Fe surface and the decomposition of PFAS with rising temperature along with the release of F – ions in the solution phase above 250 °C [125]. Thus, zero-valent Fe powder with a large specific surface area should be used in subcritical water treatment to improve the decomposition of PFAS. Supercritical water treatment is more efficient than subcritical water treatment. For example, in other tests reported by Hori et al., these treatments resulted in the decomposition of 94.8% and 84% of PFHxA, respectively, using the same concentration of Fe (9.6 mmol) [130]. However, supercritical water treatment caused an 8-fold increase in fluoroform (CF3H) creation compared to subcritical water treatment. Thus, supercritical water treatment of PFAS may be problematic from an environmental perspective as CF3H is a greenhouse gas with high diffusivity, low viscosity, and an atmospheric lifetime of approximately 270 years [131].

Moreover, both subcritical and supercritical water treatments may encounter operational problems, such as severe reactor corrosion (caused by acids formed during the treatment, such as H2SO4) or serious plugging of the reactor caused by the salts precipitating at high temperatures and low densities. This technique is not suitable for treating PFAS in landfill leachate due to operational problems and extremely high costs [132]. Page 19/30

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Lee et al. found that microwave (MW) irradiation at 140 W and 130 °C for 8 h degraded only 3% of PFOA (with an initial concentration of 241 µM in the solution) [133]. In combination with 5 mM S2O82–, the method degraded 35% and 68% of PFOA at 60 °C and 90 °C, respectively, after 4 h. The process can be further improved by adding more S2O82–. For example, in the study, the decomposition rate constant increased from 0.74 to 0.88 L/h at 90 °C when the concentration of S2O82– increased from 5 to 10 mM at acidic pH (e.g., 2.5) [133, 134]. The degradation rate of PFOA can be further increased by adding zero-valent iron; for example, Lee et al. recorded 58% degradation of PFOA in the presence of 3.6 mM Fe(0) and 5 mM S 2O82– after 1 h of MW irradiation at 90 °C [133]. Aerobic and anaerobic corrosion of Fe(0) generate Fe2+, which accelerates the decomposition of S2O82– by acting as a transitional metal activator [135]. It is important to determine the optimum concentration of Fe(0) as surplus Fe(0) may release large quantities of Fe 2+ ions that consume SO4•–, thereby reducing the decomposition efficiency of PFAS [133, 136]. Moreover, chloride ions may reduce the decomposition of PFAS as they react with free sulfate radicals that limit the oxidation efficiency [134].

Thus, degradation of PFAS by using MW irradiation in combination with oxidants may be an option for treating contaminated leachate with low buffer capacity (since pH should be adjusted to acidic values) and low concentrations of Cl –. Busch et al. [1] and Silva et al. [137] reported that PFAS treatment using wet air oxidation (WAO) was ineffective in removing PFAS from landfill leachate, most likely due to a short treatment time [1, 137] (Table 3). High concentrations of PFAS in leachates have been observed after treatment with WAO (Table 3), which can be explained by the degradation of fluorotelomer precursors to PFCAs [138]. 8. Conclusions The Registration, Evaluation, Authorization, and Restriction of Chemicals (REACH) should focus not only on long-chain but also on short-chain PFAS (such as PFBA and PFBS) because of their persistence, high mobility in water and soil, and difficulty in remediation and water purification. The use of PFAS in products should also be better regulated to reduce their release into the environment. Directives on the definition and classification of waste should be more specific about the concentrations of PFAS in wastes, such as contaminated soil. The Landfill Directive should include the maximal concentrations of soluble PFASs in waste that can be disposed of in landfills.

The introduction of legal limits for soluble/mobile PFAS in contaminated soil would reduce the amount of PFAS in landfill leachate. Additional requirements to treat the soil before landfilling and improvements in the treatment techniques would further reduce the release of PFAS into the surrounding environment. “Digging and dumping” continues to be the most widely used remediation technique at the sites contaminated with PFAS as the current in-situ and on-site remediation techniques are not reliable. Excavated soil may contain high amounts of PFAS and may comprise one of the major waste streams at landfills in the future. These soils have the potential to contaminate landfill leachate with PFAS, which places new demands on the leachate treatment systems at landfills. PFAS-contaminated soil, therefore, requires treatment before landfilling to reduce its potential to contaminate landfill leachate. The stabilization of PFAS-contaminated soil with AC, compost, Fe-oxides, or clay minerals may be an option to reduce the leaching of PFAS from landfilled soil. The potential to treat PFAScontaminated soil with another waste (e.g., carbon-rich waste) should be studied as this may reduce Page 20/30

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treatment costs, thereby enhancing sustainability. More research on the longevity of stabilized contaminants in the soil at landfills is also needed. The installation of a cover (or encapsulation) decreases the volume of leachate in contact with the PFAS-contaminated soil in a landfill. PFAS-contaminated soil should not be disposed of in a landfill together with alkaline materials to avoid increasing the mobility of PFAS. Waste in a landfill may act as a filter for leached PFAS. Furthermore, co-contaminants, including minerals and natural organic matter, can significantly affect the remediation. Thus, the transport of PFAS through landfills should be studied in more detail. Biological methods are not suitable for the treatment of PFAS in landfill leachate as they are inadequate in the reduction of PFAS concentrations. Membrane-based treatments, in combination with sorption on AC, might be viable for treating PFAS in landfill leachate. Advanced oxidation and thermal treatments also show promising results for degrading PFAS. Any treatment process selected should be optimized for, inter alia, leachate flow, treatment time, and concentrations of PFAS in untreated leachate, as well as, the desired levels of PFAS in the treated leachate. Foam fractionation is another technique that is not well-documented in the scientific literature.

Additional research is needed to study the transformation of PFAS from contaminated water to foam, the efficiency of the process, and the treatment costs. Acknowledgments This work was supported by Ragn-Sells Treatment & Detox AB and the Knowledge Foundation (KKS) within the Enforce Research Project (20160019). Author Contributions Igor Travar: Conceptualization, writing, and reviewing. Jean Noel Uwayezu: writing and reviewing. Jurate Kumpiene: reviewing and editing. Leo W.Y. Yeung: reviewing and editing. Competing Interests The authors have declared that no competing interests exist. References 1.

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Current Pollution Reports (2021) 7:524–539 https://doi.org/10.1007/s40726-021-00199-z

LAND POLLUTION (GM HETTIARACHCHI AND A JUHASZ, SECTION EDITORS)

A Review of Immobilisation‑Based Remediation of Per‑ and Poly‑Fluoroalkyl Substances (PFAS) in Soils Julie A. Sleep1 · Albert L. Juhasz1 Accepted: 1 July 2021 / Published online: 28 September 2021 © The Author(s), under exclusive licence to Springer Nature Switzerland AG 2021

Abstract Purpose of Review This review provides an overview of the latest developments in immobilisation of per- and poly-fluoroalkyl substances (PFAS) for soil remediation. It examines the efficacy of a range of amendments, including those with binding agents, along with a discussion of immobilisation mechanisms and post-immobilisation assessment needs. Recent Findings Researchers have recently applied a variety of soil amendments to soil for PFAS immobilisation. Efficacy of these has varied widely, both between amendment and soil types and for individual PFAS molecules present in contaminated soils. Activated carbon based amendments, including composite amendments exhibit the highest efficacies of the examined studies. Summary Immobilisation of PFAS is complex, with efficacy of immobilisation varying with soil properties including pH, clay and organic matter content, amendment properties, and molecular properties of the individual PFAS. Optimal remediation strategies need to be adjusted accordingly to site specific soil properties and contamination profiles. Additionally, bioavailability testing needs to supplement standard leaching approaches to determine effectiveness of PFAS soil immobilisation strategies. Keywords PFAS · Soil remediation · Immobilisation · Soil amendments · Leachability

Introduction Per- and poly-fluoroalkyl substances (PFAS) are a group of man-made chemicals found in fire-fighting foams, water repellent coatings, and consumer products such as carpets, textiles and food packaging [1, 2]. In recent years, PFAS have emerged as contaminants of concern due to their potential for persistence, mobility and bioaccumulation in humans and the environment. PFAS contamination is complex, with PFAS consisting of thousands of different compounds (over 3,000 of these commercially available) [3] although only a limited number are routinely assessed in environmental samples. Of these, perfluorooctan`e sulfonic acid (PFOS), perfluorooctanoic acid (PFOA) and perfluorooctane sulfonyl fluoride (PFOSF) are listed in the Stockholm convention, and perfluorohexane sulfonic acid (PFHxS) is listed as This article is part of the Topical Collection on Land Pollution * Albert L. Juhasz albert.juhasz@unisa.edu.au 1