Open AccessReview Advanced Degradation and Remediation Strategies for Per- and Polyfluoroalkyl Substances (PFASs): Challenges and Future Perspectives 1 Engineering Research Center of Coal-Based Ecological Carbon Sequestration Technology of the Ministry of Education, Key Laboratory of Graphene Forestry Application of National Forest and Grass Administration, Shanxi Datong University, Datong 037009, China 2 Shanxi Provincial Ecological Environment Monitoring and Emergency Support Center, Shanxi Provincial Academy of Ecological Environment Sciences, Taiyuan 030024, China 3 School of Environment, Tsinghua University, Beijing 100084, China 4 Xinjiang Institute of Ecology and Geography, Chinese Academy of Sciences, Urumqi 830011, China 5 College of Resources and Environment, University of Chinese Academy of Sciences, Beijing 100049, China 6 School of Chemical Engineering and Technology, Xi’an Jiaotong University, Xi’an 710049, China * Author to whom correspondence should be addressed. Toxics 2026, 14(6), 499; https://doi.org/10.3390/toxics14060499 (registering DOI) Submission received: 25 April 2026 / Revised: 3 June 2026 / Accepted: 4 June 2026 / Published: 7 June 2026 Abstract Per- and polyfluoroalkyl substances (PFASs) are persistent aquatic contaminants whose strong C–F bonds make conventional water treatment ineffective. This review critically synthesizes recent progress in aqueous PFAS degradation through four mechanistic routes: oxidation-driven, biodegradation, reduction-driven, and nonradical processes. Rather than evaluating technologies by parent-compound disappearance alone, we compare their defluorination and mineralization capacities, matrix tolerance, byproduct risks, energy demand, operational stability, and technology readiness. Oxidative and reductive systems can promote rapid degradation or defluorination, but their performance is often constrained by radical/electron quenching, incomplete mineralization, and sensitivity to PFAS structure and water chemistry. Biodegradation and enzymatic approaches offer mild transformation pathways but remain limited by slow kinetics, narrow substrate specificity, and uncertain toxicity evolution. Nonradical and thermochemical processes show stronger potential for deep destruction, particularly in concentrated PFAS streams. Overall, electrochemical oxidation, plasma treatment, and thermal/supercritical oxidation appear closer to practical implementation for spent adsorbents, regenerants, industrial concentrates, and other high-strength wastes, whereas many photocatalytic, biological, and microdroplet systems remain laboratory-stage. Future research should prioritize integrated separation–destruction treatment trains and standardized metrics including total organic fluorine removal, fluoride release, final residual PFAS concentrations relative to regulatory thresholds, transformation-product toxicity, energy consumption, and life-cycle impacts. 1. Introduction Perfluoroalkyl substances (PFASs), recognized as persistent and bioaccumulative “forever chemicals” with significant ecotoxicity, are extensively utilized in firefighting foams, nonstick coatings, waterproof materials, and other industrial applications [ 1, 2]. These compounds undergo widespread environmental transport and accumulation and can pose serious health risks—including immune suppression, endocrine disruption, and carcinogenicity—through bioaccumulation [ 3, 4, 5]. The environmental recalcitrance of PFASs stems primarily from the high bond energy (~544 kJ·mol −1) of the C–F bond and the steric shielding effect of fluorine atoms, which confer strong resistance to conventional degradation pathways and render traditional water treatment processes largely ineffective [ 6, 7, 8]. Current water treatment technologies, including coagulation/flocculation, biological treatment, adsorption, and ion exchange, mainly transfer or concentrate PFASs rather than destroy them, resulting in secondary pollution and costly waste management [ 9, 10, 20, 21]. Even advanced oxidation processes based on •OH or O 3 have limited ability to cleave the robust C–F bonds, highlighting the urgent need for efficient destructive treatment technologies [ 22, 23]. 2. Core Degradation Mechanisms To clarify the mechanistic basis and temporal rationale of Figure 1, the development of aqueous PFAS degradation research is summarized as an overlapping evolution of major mechanistic concepts rather than a series of strictly separated historical stages. The time windows shown in Figure 1 were assigned according to a chronological mapping of representative studies and the periods during which specific degradation mechanisms became increasingly reported in the literature. Therefore, these windows should not be regarded as exact years of first discovery or as mutually exclusive stages. Instead, oxidative/photochemical, biological, reductive, and non-radical pathways have developed in parallel, with different mechanisms becoming prominent at different periods. 2.1. Oxidation-Driven Pathways A comprehensive analysis of the data in Table 1 indicates that various advanced oxidation processes (AOPs) degrade PFASs primarily through the attack of highly reactive radicals (mainly •OH and SO 4• −) on the carboxyl functional groups, triggering decarboxylation and C–F bond cleavage. The degradation efficiency is governed by both the intrinsic oxidizing power of the reactive species and their effective steady-state concentrations in the system [ 51, 58, 66]. Although the oxidation potentials of •OH (~2.8 V vs. NHE) and SO 4• − (2.5–3.1 V vs. NHE) are thermodynamically sufficient to drive C–F bond cleavage, notable differences in degradation kinetics are observed among different AOPs operating at similar potentials. Electrochemical oxidation using BDD or Ti 4O 7 anodes achieves PFOA degradation rate constants in the range of 0.019–0.1838 min −1 [ 11, 12, 53, 55, 56, 57, 58], whereas the rate constants for photocatalytic oxidation typically fall between 0.0037 and 0.1 min −1 [ 47, 48, 49, 50, 51]. This contrast clearly demonstrates that the oxidation potential is not the sole determinant of the degradation rate. The observed differences in degradation kinetics are directly associated with each system’s capacity to generate and maintain steady-state concentrations of reactive species. Plasma oxidation simultaneously produces •OH, O 3, e aq −, and reactive nitrogen species, creating a synergistic oxidative-reductive environment that enables PFOA degradation rates of 0.015–0.085 min −1, with the effective concentration maintained through gas–liquid interfacial enrichment [ 60, 61, 62, 63, 64, 65, 66]. Persulfate-based activation techniques primarily generate SO 4• −, achieving PFOA degradation rates of up to 0.1838 min −1 while maintaining high steady-state concentrations [ 11, 58, 68, 69]. Sonolytic oxidation relies on cavitation-induced localized •OH generation and extreme reaction conditions [ 14, 28]; supercritical water oxidation, on the other hand, utilizes •OH under high temperature and pressure to achieve >99.999% PFAS removal [ 13]. 2.2. Biodegradation Pathways According to dominant biological functions and degradation mechanisms, PFAS biodegradation technologies can be categorized into three main types: biosorption-assisted cometabolism, physicochemical pretreatment-enhanced biotransformation, and direct degradation by specialized defluorinating microorganisms. Biosorption-assisted cometabolism, such as algal–bacterial granular sludge and osmotic microbial fuel cells, relies on extracellular polymeric substances to concentrate PFAS through electrostatic, hydrogen-bonding, and hydrophobic interactions, while microbial metabolic pathways help maintain system activity [ 29, 76, 77]. Pretreatment-enhanced biotransformation combines abiotic activation with microbial or enzymatic degradation. For example, photocatalytic–fungal systems can weaken PFAS structures before fungal enzyme-mediated transformation, while microencapsulation can enrich PFOS and protect functional microorganisms or dehalogenases [ 16, 75]. Direct degradation by specialized microorganisms, such as Acidimicrobium sp. A6 and Pseudomonas species, has also been reported to involve reductive dehalogenase, dehalogenase, or monooxygenase-related pathways that contribute to partial defluorination and chain shortening [ 17, 19, 78]. 2.3. Reduction-Driven Pathways The hydrated electron (e aq −), possessing a highly negative reduction potential of −2.9 V, is the most effective species, with electron transfer rate constants reaching the order of 10 6–10 9 M −1 s −1 [ 80, 81, 82, 83]. Technologies based on e aq − generation, such as UV/VUV/sulfite systems and plasma reduction, exhibit relatively high defluorination kinetics, with pseudo-first-order rate constants up to 0.016–0.37 min −1 [ 87, 88, 89, 102]. Despite these kinetic advantages, the high bond dissociation energy (BDE) of the C–F bond (approximately 106.8–485 kJ/mol or ~5 eV) remains a primary thermodynamic challenge [ 80, 81, 85, 92]. In contrast, cathodic electrons in electrochemical reduction operate at a less negative potential (−1.8 to −1.4 V vs. SHE), resulting in lower electron transfer rate constants (~10 −4 s −1) and slower defluorination rates [ 90, 91]. Photocatalytic and photoelectrocatalytic systems rely on conduction band electrons (e_cb −), whose reduction potential (approximately −0.99 to −0.76 V vs. SCE) further limits the thermodynamic driving force, leading to slower degradation rates (0.2–1.53 h −1) [ 93, 95]. Several technologies enhance efficiency through multispecies synergy or interfacial effects. Plasma-cathode electrochemical reduction utilizes energetic gaseous electrons for direct C–F bond attack [ 92]. The Pd 0-biofilm synergistic system combines chemical hydrogenation by hydrogen atoms (H•) with the biocatalysis of microbial enzymes [ 96]. Synergistic processes such as microwave discharge of plasma-Fe 2+ and UV/zero-valent iron promote the generation and utilization of e aq − through the mediation of Fe 2+ [ 97, 98]. Furthermore, systems such as gas–liquid interface spontaneous reduction systems and solar-driven UV/sulfite systems have the potential for reductive defluorination at specific environmental interfaces or under natural energy driving forces [ 100]. Collectively, reductive pathways face a dual challenge: sustaining the efficient generation of e aq − against quenching by background constituents in real-water matrices while concurrently overcoming the high energy barrier inherent to C–F bond cleavage. Future research should focus on developing more efficient and stable electron donor systems, optimizing interfacial electron transfer processes, and exploring synergistic degradation mechanisms coupled with biological or oxidative processes to achieve efficient and deep defluorination of short-chain and structurally complex PFASs. 2.4. Nonradical Pathways In addition to radical-mediated mechanisms, nonradical pathways constitute another critical route for PFAS degradation ( Table 4). These pathways do not rely on diffusible radicals such as •OH or SO 4• − but instead attack PFAS molecules directly via electron transfer or thermo/chemical activation, particularly by targeting their carboxyl functional groups [ 103, 104]. The adsorption–photocatalytic synergistic technology relies primarily on photogenerated holes (h +) at semiconductor surfaces to directly oxidize PFAS. The degradation followed pseudo-first-order kinetics, with rate constants ranging from 0.08 to 0.918 h −1 [ 105, 107]. Direct electrochemical oxidation achieves degradation through direct electron transfer at the anode surface, which also results in a broad kinetic range (pseudo-first-order k = 0.0175–0.946 min −1) [ 111, 112]. Its efficiency is closely related to the anode material, applied potential, and PFAS structure. Thermochemical pathways primarily disrupt C–F and C–C bonds directly via thermal energy. Thermal decomposition technology degrades PFAS through pyrolysis at elevated temperatures, with pseudo-first-order rate constants between 0.02 and 0.14 min −1 [ 116, 117]. Hydrothermal degradation, which occurs in high-temperature and high-pressure aqueous environments, achieves PFAS transformation via the synergistic effect of OH − nucleophilic attack on the carboxyl group and catalytic decomposition, resulting in a degradation rate constant for PFOA of approximately 0.0204 min −1 [ 124]. In conclusion, while nonradical pathways are distinguished by higher reaction selectivity and robustness against quenching in complex water matrices than their radical counterparts are, their degradation efficiency is frequently constrained by kinetic bottlenecks in interfacial electron transfer or by the magnitude of the applied thermal/chemical energy. These pathways are particularly suitable for treating systems with medium to high PFAS concentrations and low mass transfer limitations or as complementary/synergistic approaches to radical processes. Future research should focus on elucidating the interfacial electron transfer and molecular-level mechanisms and on developing efficient, stable nonradical active interfaces or reaction systems to expand the technological boundaries for PFAS degradation. 3. Progress in Degradation Technology and Performance Evaluation 3.1. Radical Oxidation Technologies The comparative analysis in Table 5 shows that radical-dominated advanced oxidation technologies (AOTs) can achieve more than 90% removal of PFAS, such as PFOA and PFOS, under optimized laboratory conditions. However, defluorination efficiencies vary widely from 33% to 90%, indicating that parent-compound removal does not necessarily represent complete fluorine mineralization [ 14]. Most studies remain at the laboratory scale, although electrochemical oxidation using Ti 4O 7 anodes and plasma oxidation have shown pilot or transitional-scale potential, maintaining >90% PFAS removal across concentrations from µg/L to mg/L [ 13]. Nevertheless, their performance toward short-chain and emerging PFASs, long-term stability in complex real-water matrices, energy consumption, and overall cost-effectiveness remain insufficiently demonstrated [ 11, 12]. Table 5. Summary and Comparison of PFAS Degradation Performance by Radical-based Oxidation Technologies. Table 5. Summary and Comparison of PFAS Degradation Performance by Radical-based Oxidation Technologies. Technology Target PFAS Scale Initial Conc. (mg/L) Reaction Conditions Time (h) Degradation (%) Defluorination (%) Refs. 3.2. Biodegradation Technologies Biodegradation technologies have attracted interest because they may operate under mild conditions and require lower external energy input than many physicochemical processes. However, PFAS biodegradation remains at an early and highly constrained stage of development. These processes depend on microbial enzyme systems, such as dehalogenases and monooxygenases, and their performance is strongly affected by PFAS structure, microbial community composition, enzyme expression, redox conditions, cosubstrate availability, and reaction time [ 16, 17, 18, 19, 29, 75, 76, 77, 78]. As shown in Table 6, biosorption- and cometabolism-based systems, such as algal–bacterial granular sludge and osmotic microbial fuel cells, can achieve 40–60% PFAS removal but require long reaction cycles and show limited direct C–F bond cleavage [ 29]. In contrast, enhanced biotransformation and specialized microbial degradation pathways exhibit greater defluorination potential; for example, photocatalytic–fungal coupling achieved 90% PFOA degradation and 60% defluorination, while Acidimicrobium sp. A6 achieved 60% degradation and 80% defluorination under extended anaerobic cultivation [ 17, 75]. Table 6. Efficiency and Mechanisms of PFAS Degradation via Biological Pathways. Table 6. Efficiency and Mechanisms of PFAS Degradation via Biological Pathways. Technology Target PFAS Scale Initial Conc. (mg/L) Reaction Conditions Time (h) Degradation (%) Defluorination (%) Refs. 3.3. Reductive Degradation Technologies Reductive pathways based on hydrated electrons (e aq −) have shown strong PFAS defluorination potential under laboratory conditions. As summarized in Table 7, sulfite photoreduction and plasma reduction can achieve >90% degradation and 58.9–98% defluorination for various long-chain PFASs [ 59, 65, 80, 82, 83, 85, 87]. Synergistic systems, such as DBD/sulfite–ultrafiltration coupling, can further increase PFOA defluorination to 98%, indicating the performance-enhancing effect of multiprocess integration [ 101]. Table 7. Performance Evaluation of Reductive Technologies for PFAS Treatment. Table 7. Performance Evaluation of Reductive Technologies for PFAS Treatment. Technology Target PFAS Scale Initial Conc. (mg/L) Reaction Conditions Time (h) Degradation (%) Defluorination (%) Refs. Sulfite Photoreduction PFOS Bench 10 UV/VUV, 10 mM SO 32−, 22 °C 6 64 58.9 [ 80] PFHxS 52.6 36.8 6:2 FTSA 37.8 28.4 PFBS UV/VUV, 10 mM SO 32−, 10 mM KI, 85 °C 38 18 PFOA Bench 0.1 pH = 12, 10 mM SO 32−6 100 66.2 [ 81] PFBA 99.6 47.8 GenX 70 42.3 PFHpA Bench 5.1 VUV/UV, 10 mM SO 32−, subsequent oxidation 8 83.5 74 [ 83] PFHxS 5.2 8 74 73.4 6:2 FTS 5.4 12 98.9 86.9 PFOA Bench 5 UV, 1.5 mg/L BrO 3−4 96 96 [ 82] PFOS 90 90 PFOA Bench 36 μM UV, SO 32−12 100 73.8 [ 84] PFOS 32 μM 0.5 98 55 GenX 0.16 mM 6 100 90 PFOS 1.2 mM Visible light 405 nm, CdS NCs 8 100 100 PFOA Bench 36 μM VUV, 20 μM Fe 3+4 51.2 51.2 [ 85] PFOS 32 μM UV, SO 32−0.5 98 70 PFOA Bench 1 pH = 10, UV irradiation 1 100 98 [ 71] Plasma Reduction F-53B Bench 0.5 Discharge power 94 W, gas–liquid interface 0.33 100 ~45 [ 89] PFOS 10 1.33 89.5 48.3 PFOA Bench 0.2 N 2 plasma, 60 min 1 45 - [ 126] PFDA 0.2 N 2 plasma, 60 min 1 60 - PFDA Bench 0.0062 Bipolar discharge, Hyamine 1622, air 1.25 >97 82 [ 86] PFNA 0.0118 >97 82 PFOS 0.0087 >97 82 PFOA 0.0163 >97 82 PFHpA 0.0139 >97 82 PFBS 0.0191 95 82 PFBA 0.0103 53 82 PFOS Pilot 0.1365 Fine bubble, 2 J/pulse, Ar 1 L/min 2 >99.5 - [ 102] PFOS Bench 20 20 kV, 50 ns, 1 kHz, Ar 1 100 - [ 127] PFOS Bench 10 30 W, Bi 2O 3 catalyst, 25 °C 0.5 96.6 68.6 [ 87] PFOA Bench 41.4 Ar bubbling, 4 W input, room temp. 0.5 98.9 64.4 [ 59] PFOS 5 0.17 >99 - PFBS Bench 10 Gas–liquid discharge, 0.2 mM C 12TAB, 10 °C, Ar 0.5 100 - [ 90] PFOA Pilot 8.3 Ar 4 L/min, 40 Hz, −30 kV, room temp. 1 90 - [ 65] PFOS 1 90 - Electrochemical Reduction PFMeUPA Bench 49.5 10 V, 70 °C, 0.1 M Na 2SO 45 100 ~90 [ 90] PFOS 50 15 V, 85 °C, 0.1 M Na 2SO 4, 250 μM B12 4 7.9 7.9 GenX Bench 5 BDD anode + Au cathode, 0.1 M Na 2SO 4+NaCl 2 92 - [ 54] PFOA Bench 413 pH = 9.5, 0.5 M KHCO 3, Au electrode, −1.80 V 14 100 ~35 [ 91] Plasma-Cathode Electrochemical Reduction PFOA Bench 6.2 Plasma cathode, 2 mA, Ar purge 2 100 ~60 [ 92] Photocatalytic/ Photoelectrocatalytic Reduction PFOS Bench 40 365 nm LED, TEOA 24 99 97 [ 93] PFOA Bench 10 UV, pH = 6, ZnS-40% [N] 3.5 >90 >80 [ 94] PFOS 3 >90 85 PFHxA >90 70 6:3-FTCA >90 60 HFPO-TA 90 30 PFBA 90 40 PFOA Bench 0.5 Simulated sunlight, PMS activation, pH 3–9 0.67 93.6 56.7 [ 95] Pd 0-Biofilm Synergistic Reduction PFOA Bench 0.0041 H 2 MBfR, pH = 7.2, co-existing nitrate 24 85 81 [ 96] Microwave Discharge Plasma-Fe 2+Synergistic Reduction PFOA Bench 40 150 W, pH = 3.8, 10 mg/L Fe 2+1.67 98.6 82.9 [ 97] UV/Zero-Valent Iron Photocatalytic Reduction PFNA Bench 0.0005 UV 254 nm, 100 mg/L Fe 0, pH = 3 2 90 - [ 98] PFOS 88 - PFOA 46 - UV/DIHA Heterogeneous Photoreduction PFOA Bench 10 UV 254 nm, DIHA nanospheres, pH 4–10 12 100 79 [ 99] PFOS 100 79 6:3-FTCA 100 66 HFPO-DA 100 45 PFBS 51 51 TFA 9 9 Gas–Liquid Interface Spontaneous Reduction PFOA Bench 0.02 Microdroplet, pH = 9.8, 0.1 mM NaBr, air Short 36.4 36.4 [ 100] Solar-Driven UV/Sulfite Reduction PFOA Pilot 1 pH = 10, solar UV irradiation 24 100 89 [ 15] DBD/Sulfite–Ultrafiltration Coupled Process PFOA Bench/Pilot 0.1 Sulfite 2 mM, discharge power 80 W, Ar 2 L/min 1.5 87.9 98 [ 101] 3.4. Nonradical Pathways Technologies Table 8. Summary of PFAS Degradation Efficiency through Non-radical Pathways. Table 8. Summary of PFAS Degradation Efficiency through Non-radical Pathways. Technology Target PFAS Scale Initial Conc. (mg/L) Reaction Conditions Time (h) Degradation (%) Defluorination (%) Refs. To further relate the degradation efficiencies in Table 8 to regulatory relevance, it should be noted that PFAS limits vary substantially among jurisdictions and are generally defined at the ng/L level for drinking water. For example, representative values include the U.S. EPA maximum contaminant levels of 4 ng/L for PFOA and PFOS and 10 ng/L for PFHxS, PFNA, and HFPO-DA [ 128], the EU Drinking Water Directive values of 0.1 µg/L for the Sum of PFAS and 0.5 µg/L for PFAS Total [ 129], the Health Canada objective of 30 ng/L for the sum of 25 specified PFAS [ 130], and the Australian guideline values of 200 ng/L for PFOA, 8 ng/L for PFOS, 30 ng/L for PFHxS, and 1000 ng/L for PFBS [ 131]. Therefore, regulatory compliance cannot be determined from degradation or defluorination percentages alone, but should be assessed using measured final concentrations, analytical detection limits, and the specific PFAS compounds included in each regulatory framework. When the initial concentration and degradation efficiency reported in Table 8 are used to estimate the residual parent-PFAS concentration as residual = C0 × (1 − degradation/100), most high-concentration bench-scale studies remain difficult to interpret from a compliance perspective. For instance, even >99% degradation of PFOA from an initial concentration of 100 mg/L could still correspond to a residual concentration below 1 mg/L, which is far above ng/L-level drinking-water criteria. Similarly, entries reported as “100%” degradation cannot be considered compliant unless the final concentration and method detection limit are provided. In contrast, the low-initial-concentration thermal decomposition entry for total PFAS (0.00041 mg/L, 99.89% degradation) gives an estimated residual concentration of approximately 0.45 ng/L, and the hydrothermal degradation entry for PFOA (0.0002 mg/L, >99% degradation) gives an estimated residual concentration below 2 ng/L. These values are below the representative regulatory limits listed above, subject to confirmation by measured final concentrations and the applicable regulatory target analytes. Thus, Table 8 demonstrates the destruction potential of nonradical and thermochemical processes, but regulatory compliance should be claimed only when final PFAS concentrations, fluorine mass balance, and transformation-product toxicity are adequately reported. 3.5. Critical Synthesis: Technology Readiness and Realistic Application Windows Although the technologies summarized above have demonstrated potential for PFAS degradation, their practical values differ substantially. High parent-compound removal under optimized laboratory conditions does not necessarily indicate engineering readiness, because incomplete defluorination, high energy input, poor matrix tolerance, limited material durability, and the formation of short-chain or partially fluorinated byproducts may restrict field application [ 11, 12, 13, 14]. Therefore, practical applicability should be assessed using multiple criteria, including defluorination and mineralization efficiency, validation scale, operational stability in real-water matrices, energy and chemical consumption, byproduct control, and secondary-waste management [ 9, 10, 13, 17, 29]. Based on these criteria, Table 9 classifies the reviewed technologies according to their approximate technology readiness, energy demand, relative cost, realistic application windows, and major barriers. Adsorption and ion exchange should be considered as supporting separation technologies because PFAS contamination in raw water, groundwater, and drinking-water sources often occurs at low concentrations. These mature technologies can capture and concentrate PFASs, particularly long-chain compounds, from large volumes of dilute water [ 9, 10]. However, they mainly transfer PFASs to adsorbents or regenerants rather than cleaving C–F bonds. Thus, their practical role is better defined as front-end separation and enrichment, followed by downstream destructive treatment of spent media, regenerants, or concentrates. 4. Critical Factors and Challenges in the Translation of Technology from the Laboratory to Practice Based on the TRL-based application-oriented assessment in Table 9, the key challenge in PFAS remediation has shifted from demonstrating degradation mechanisms under controlled laboratory conditions to developing robust, economical, and scalable treatment systems. Degradation efficiency alone cannot determine practical applicability, because technologies with high PFAS removal may still be limited by high energy demand, costly electrodes or reactors, strict matrix requirements, insufficient scale-up validation, incomplete defluorination, byproduct formation, and secondary-waste management. Therefore, practical evaluation should consider technology readiness, energy demand, cost, long-term stability, matrix tolerance, byproduct control, and waste management. 4.1. Influence of the Pollutant Molecular Structure The molecular structure of PFAS is the intrinsic determinant of its degradation behavior, occupying a central position in the integrated challenge network shown in Figure 3. The carbon chain length directly influences the hydrophobicity and interfacial adsorption propensity. Typically, long-chain PFASs (e.g., PFOA and PFOS) exhibit shorter degradation half-lives, whereas short-chain analogs (e.g., PFBA and TFA) are more difficult to remove effectively because of their higher water solubility and mobility [ 80, 112]. The chemical nature of functional groups (e.g., sulfonate vs. carboxylate) further influences the interaction mechanisms with reactive species or catalytic material surfaces by modulating the molecular charge distribution and spatial configurations [ 81, 113]. Additionally, the steric hindrance introduced by branched structures generally slows degradation kinetics, while relatively weaker chemical bonds within the molecule (e.g., ether bonds, C–H bonds) often become key sites for initiating degradation chain reactions, governing the overall reaction pathways and rates [ 81, 113]. 4.2. Regulatory Role of Operational Parameters in Process Performance The operational parameters of a degradation system are key external variables for controlling the reaction kinetics and selectivity. Their optimization pathways and strategies run through the decision-making process from the laboratory to scale-up applications, as outlined in Figure 2. The solution pH not only affects the ionic state of PFAS but also profoundly affects the stability and generation efficiency of reactive species (e.g., hydrated electrons, e aq −; holes, h +) [ 80, 81]. In thermochemical processes, temperature directly influences the activation energy, with PFASs of different structures having distinct pyrolysis thresholds [ 119, 120]. For technologies driven by external fields such as electrochemistry and plasma, the current density and discharge power require fine-tuning to maximize the yield of active species while suppressing undesired side reactions triggered by energy overload [ 113]. 4.3. Constraining Effects of the Environmental Matrix on Process Efficiency 4.4. Pivotal Role of Material Properties in Determining Performance and Functionality 4.5. Technoeconomic Assessment and Optimization for Sustainable Implementation Promoting the engineering application of degradation technologies necessitates systematic technoeconomic analysis, a process directly corresponding to the node “Assessment of Economic Feasibility and Sustainability” in Figure 2. The specific energy consumption varies significantly among different technologies, ranging from approximately 0.16 kWh/m 3 for highly efficient coupled systems to over 1600 kWh/m 3 for some physical separation processes [ 120]. Energy reduction strategies include developing low-energy consumption combined processes (e.g., plasma–biological coupling) and optimizing energy input modes and reactor design [ 61]. Moreover, the long-term stability and regenerability of materials are critical for determining operational costs. Materials with good cyclic performance and simple regeneration methods hold greater potential for large-scale applications. 4.6. Control of Degradation Pathways and Associated Byproduct Risks 5. Conclusions and Future Perspectives The effective degradation and deep mineralization of per- and polyfluoroalkyl substances (PFASs) depend on overcoming the exceptional stability of C–F bonds. This review summarizes four major degradation pathways: oxidation, biodegradation, reduction, and nonradical processes. Oxidation technologies enable rapid degradation but are vulnerable to matrix-induced radical quenching and material costs. Biodegradation offers a potentially mild route for PFAS transformation, but its current applicability is limited by slow kinetics, substrate specificity, microbial and redox dependence, and uncertain intermediate toxicity, making it more suitable as a supporting or polishing step than as a stand-alone solution. Reductive technologies based on hydrated electrons show strong defluorination potential, especially for long-chain PFASs, but remain sensitive to PFAS structure and water-matrix interference. Nonradical and thermochemical pathways can achieve deep destruction under suitable conditions, particularly for concentrated PFAS wastes. A key conclusion is that PFAS degradation technologies should not be evaluated solely by parent-compound removal under optimized laboratory conditions. Practical assessment should also consider defluorination and mineralization efficiency, technology readiness, energy demand, cost, matrix tolerance, material durability, byproduct control, secondary-waste management, and long-term operation. Electrochemical oxidation, plasma treatment, and thermal/supercritical oxidation are relatively closer to practical implementation for concentrated PFAS streams, spent adsorbents, ion-exchange regenerants, foam fractions, and industrial concentrates. In contrast, UV/VUV/sulfite photoreduction, sonochemical degradation, persulfate activation, and selected nonradical catalytic systems require further continuous-flow and real-water validation, whereas biodegradation, enzymatic degradation, microdroplet reactions, and many photocatalytic systems remain mainly at the laboratory or proof-of-concept stage. PFAS treatment strategies should be matched to contamination scenarios. For low-concentration raw water, groundwater, and drinking-water sources, adsorption and ion exchange are more realistic front-end separation and enrichment technologies, but downstream destructive treatment of spent media, regenerants, or concentrates is still required. For high-strength PFAS wastes, direct destructive technologies such as electrochemical oxidation, plasma treatment, hydrothermal treatment, and supercritical water oxidation may be more appropriate. From a toxicity- and compliance-oriented perspective, PFAS remediation should not be judged only by parent-compound disappearance, carbon-chain shortening, or percentage removal. Future studies should include final residual PFAS concentrations, analytical detection limits, fluorine mass balance, fluoride recovery, total organic fluorine reduction, transformation-product identification and toxicity, and comparison with jurisdiction-specific regulatory thresholds to better confirm practical compliance and genuine environmental and health risk reduction. Future research should advance synergistically across the following fronts: (1) Materials and reactor innovation. Durable, low-cost, and selective electrodes, catalysts, membranes, and functional materials should be developed to improve PFAS degradation efficiency, electron or radical utilization, matrix resistance, and long-term stability. Reactor design should also be optimized to enhance mass transfer, energy efficiency, scalability, and continuous operation. (2) Scenario-oriented technology integration. Treatment systems should be designed according to PFAS concentration, matrix composition, and waste-stream characteristics. For dilute waters, an “enrichment–destruction–polishing” strategy is more realistic, whereas concentrated PFAS residuals may be more suitable for direct destructive treatment. Coupling separation with electrochemical, plasma, thermal, reductive, or biological polishing processes may improve both efficiency and economic feasibility. (3) Real-water and engineering-scale validation. Future studies should move beyond simplified laboratory matrices and evaluate long-term performance in representative real waters containing dissolved organic matter, inorganic ions, salinity, surfactants, suspended solids, and coexisting micropollutants. Pilot-scale and continuous-flow validation, together with standardized metrics such as energy consumption per unit PFAS destroyed, fluoride recovery, material lifetime, treatment cost, and life-cycle impacts, is needed to bridge laboratory performance and field application. (4) Pathway control and risk management. Greater attention should be paid to degradation pathways, intermediate evolution, fluorine mass balance, and byproduct toxicity. Reaction conditions should be regulated to promote deep defluorination and mineralization while minimizing persistent short-chain or partially fluorinated products. Comprehensive monitoring is needed to ensure genuine risk reduction. Overall, no single technology can currently serve as a universal solution for PFAS-contaminated water. Future progress will depend on matching mechanisms with realistic application scenarios, integrating separation and destruction processes, validating performance under complex real-water conditions, and evaluating treatment outcomes using both chemical and toxicity-oriented indicators. Author Contributions Conceptualization, X.Z., W.S. and L.P.; data curation, T.H., R.Z. and X.Y.; formal analysis, T.H. and R.Z.; investigation, R.Z. and W.S.; project administration, X.Y. and J.Z.; supervision, P.D. and L.P.; visualization, T.H. and R.Z.; writing—original draft, X.Z. and T.H.; writing—review & editing, X.Z., J.Z., L.P. and P.D. All authors have read and agreed to the published version of the manuscript. Funding This research received no external funding. Institutional Review Board Statement This study does not involve humans and animals. Informed Consent Statement This study does not involve any humans. Data Availability Statement No new data were created or analyzed in this study. Data sharing is not applicable to this article. Acknowledgments The authors would like to acknowledge the kind help and suggestions of all the anonymous reviewers. Conflicts of Interest The authors declare no conflicts of interest. References Figure 1. Approximate evolution of major mechanistic concepts and representative technology groups in aqueous PFAS degradation. The time windows indicate periods during which each mechanistic category became increasingly reported in the literature, rather than exact years of first discovery. The categories overlap because oxidative/photochemical, biological, reductive, and non-radical pathways have developed in parallel. Representative technologies and dominant mechanisms are shown to clarify the basis for classifying PFAS degradation technologies. Figure 1. Approximate evolution of major mechanistic concepts and representative technology groups in aqueous PFAS degradation. The time windows indicate periods during which each mechanistic category became increasingly reported in the literature, rather than exact years of first discovery. The categories overlap because oxidative/photochemical, biological, reductive, and non-radical pathways have developed in parallel. Representative technologies and dominant mechanisms are shown to clarify the basis for classifying PFAS degradation technologies. Figure 2. Decision Pathway from Laboratory to Field-Scale Application. Figure 2. Decision Pathway from Laboratory to Field-Scale Application. Figure 3. Network Diagram of Integrated Challenges Facing PFAS Degradation Technology Application. Figure 3. Network Diagram of Integrated Challenges Facing PFAS Degradation Technology Application. Table 1. Mechanisms and Kinetic Characteristics of Active Species Formation in PFAS Advanced Oxidation Processes. Table 1. Mechanisms and Kinetic Characteristics of Active Species Formation in PFAS Advanced Oxidation Processes. Technology Key Reactive Species Oxidation Potential Rate Constant with Functional Group Steady-State Concentration of Reactive Species Refs. Table 2. Key Enzyme Systems and Gene Expression Profiles for the Biocatalytic Degradation of PFAS. ↑ indicates an increase or upregulation compared with the corresponding control or baseline condition. Table 2. Key Enzyme Systems and Gene Expression Profiles for the Biocatalytic Degradation of PFAS. ↑ indicates an increase or upregulation compared with the corresponding control or baseline condition. Technology Key Active Species Changes in Functional Gene Expression Enzyme–Substrate Interaction Mechanism Refs. Table 3. Characteristics and Performance Comparison of PFAS Reductive Defluorination via Hydrated Electrons. ΔG‡ denotes the activation Gibbs free energy, and ‡ indicates the transition state. Table 3. Characteristics and Performance Comparison of PFAS Reductive Defluorination via Hydrated Electrons. ΔG‡ denotes the activation Gibbs free energy, and ‡ indicates the transition state. Technology Key Reactive Species Reduction Potential Electron Transfer Rate Constant Defluorination Energy Barrier Refs. Table 4. Mechanisms and Kinetics of PFAS Degradation by Nonradical Pathways. Table 4. Mechanisms and Kinetics of PFAS Degradation by Nonradical Pathways. Technology Key Active Species Direct Electron Transfer/Degradation Kinetics Refs. Adsorption–Photocatalytic Synergy Surface Holes (h +) Pseudo-first-order rate constant (k) = 0.08–0.918 h −1[ 105, 106, 107, 108, 109, 110] Electrochemical Direct Oxidation Direct Electron Transfer Pseudo-first-order k = 0.0175–0.946 min −1[ 111, 112, 113, 114, 115] Thermal Decomposition Pyrolytic Action Pseudo-first-order k = 0.02–0.14 min −1[ 116, 117, 118, 119, 120, 121, 122] Hydrothermal Degradation Catalytic Decomposition + OH − Nucleophilic Attack Pseudo-first-order k = 0.0204 min −1 (for PFOA) [ 123, 124, 125] Table 9. TRL-based application-oriented assessment of PFAS degradation and supporting treatment technologies. Table 9. TRL-based application-oriented assessment of PFAS degradation and supporting treatment technologies. Technology Representative Technologies Practical Readiness Energy Demand Relative Cost Most Realistic Application Window Main Barriers Refs. Relatively close to implementation Electrochemical oxidation, plasma treatment, thermal decomposition, supercritical water oxidation Medium to high High High Concentrated PFAS streams, spent adsorbents, regenerants, and industrial wastewater High energy demand, electrode/reactor cost, long-term stability, mass-transfer limitation, byproduct control [ 11, 12, 13, 52, 53, 54, 55, 56, 57, 58, 59, 60, 61, 62, 63, 64, 65, 66, 111, 112, 113, 114, 115, 116, 117, 118, 119, 120, 121, 122, 123, 124, 125] Transitional technologies requiring pilot validation UV/VUV/sulfite reduction, persulfate activation, sonochemical degradation, selected nonradical catalytic or electrochemical systems Medium Medium to high Medium to high Pretreated waters, concentrated streams, and hybrid treatment systems Matrix quenching, reagent consumption, incomplete defluorination, limited continuous-flow and real-water data [ 14, 15, 28, 69, 70, 71, 80, 81, 82, 83, 84, 85, 93, 94, 95, 105, 106, 107, 108, 109, 110, 111, 112, 113, 114, 115] Mainly laboratory-stage technologies Biodegradation, enzymatic degradation, microdroplet reactions, many photocatalytic/photoelectrocatalytic systems Low to medium Low to medium Low to medium at bench scale; uncertain at scale-up Polishing steps, pretreatment-assisted systems, and niche applications Slow kinetics, substrate specificity, idealized reaction conditions, limited mineralization evidence [ 16, 17, 18, 19, 29, 47, 48, 49, 50, 51, 72, 73, 74, 75, 76, 77, 78, 93, 94, 95, 105, 106, 107, 108, 109, 110] Supporting/enabling separation technologies Adsorption and ion exchange High for separation, low for destruction alone Low to medium Medium; depends on media regeneration and disposal Front-end enrichment before destructive treatment Phase transfer rather than destruction, secondary waste generation, spent media disposal [ 9, 10] Disclaimer/Publisher’s Note: The statements, opinions and data contained in all publications are solely those of the individual author(s) and contributor(s) and not of MDPI and/or the editor(s). MDPI and/or the editor(s) disclaim responsibility for any injury to people or property resulting from any ideas, methods, instructions or products referred to in the content. © 2026 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license. Share and Cite MDPI and ACS Style Zhang, X.; Han, T.; Yao, X.; Zhao, R.; Sun, W.; Pei, L.; Zhao, J.; Duan, P. Advanced Degradation and Remediation Strategies for Per- and Polyfluoroalkyl Substances (PFASs): Challenges and Future Perspectives. Toxics 2026, 14, 499. https://doi.org/10.3390/toxics14060499 AMA Style Zhang X, Han T, Yao X, Zhao R, Sun W, Pei L, Zhao J, Duan P. Advanced Degradation and Remediation Strategies for Per- and Polyfluoroalkyl Substances (PFASs): Challenges and Future Perspectives. Toxics. 2026; 14(6):499. https://doi.org/10.3390/toxics14060499 Chicago/Turabian Style Zhang, Xiaohui, Tongshun Han, Xiaofeng Yao, Rui Zhao, Wenjun Sun, Liang Pei, Jianguo Zhao, and Peigao Duan. 2026. "Advanced Degradation and Remediation Strategies for Per- and Polyfluoroalkyl Substances (PFASs): Challenges and Future Perspectives" Toxics 14, no. 6: 499. https://doi.org/10.3390/toxics14060499 APA Style Zhang, X., Han, T., Yao, X., Zhao, R., Sun, W., Pei, L., Zhao, J., & Duan, P. (2026). Advanced Degradation and Remediation Strategies for Per- and Polyfluoroalkyl Substances (PFASs): Challenges and Future Perspectives. Toxics, 14(6), 499. https://doi.org/10.3390/toxics14060499 Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details . Article Metrics Article metric data becomes available approximately 24 hours after publication online.
