Insights into Photo/Electrocatalysts for the Degradation of Per- and Polyfluoroalkyl Substances (PFAS) by Advanced Oxidation Processes

Abstract

Per- and polyfluoroalkyl substances (PFASs) are an emerging group of persistent organic pollutants in aquatic environments with high levels of toxicity and bioaccumulation. The risks posed by PFASs to the environment and health have attracted increasing attention. To remove them from water, advanced oxidation processes (AOPs), with the merits of high efficiency and low cost, are mainly used. Photo/electrocatalytic heterogeneous AOPs, with the assistance of nanostructured catalysts and external energy in the form of light/electricity, have emerged as one of the most powerful techniques, overcoming the difficulty associated with defluorination and achieving the effective and complete degradation of PFASs in water. The structures of photo/electrocatalysts play a critical role in the production of reactive oxygen species, the electron transfer process, and the degradation pathway and its efficiency. Herein, to elucidate the structure–performance relationship, a review of photo/electrocatalysts for the enhanced degradation of PFASs in heterogeneous AOPs, organized according to their composition and nanostructure design, is provided. This review article is mainly focused on (1) the mechanisms and pathways of PFAS degradation by heterogeneous photo/electrocatalytic AOPs, and (2) the structural designs and modifications of photo/electrocatalysts for the enhanced degradation of PFASs by heterogeneous AOPs. Finally, the challenges and prospects for future research into photo/electrocatalysts of heterogeneous AOPs in the field of PFAS remediation are discussed.

1. Introduction

Per- and polyfluoroalkyl substances (PFASs) are an emerging group of persistent organic pollutants in the environment which pose ecological and health risks [1]. PFASs are man-made chemicals that are widely used in various industrial and commercial products; their chemical structure includes a fully fluorinated carbon chain with a terminated functional group attached to it. The most common terminal groups are carboxylic acid (−COOH) and sulfonic acid (−SO₃H) groups, while the fluorinated carbon chain varies in length and number of branches, containing a number of carbon–fluorine bonds. Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS), with structural formulas shown in Figure 1a, are the two most commonly detected PFASs. The carbon–fluorine bond in these molecules is one of the strongest single bonds due to its bond energy of 485–582 kJ/mol and its redox potential of F/F⁻ at 3.6 eV, making it difficult for PFASs to break down naturally [2]. Moreover, the fluorine atoms in PFASs provide a shielding effect that protects the carbon–fluorine bond from chemical and biological attack, further contributing to the persistence of these compounds. On the other hand, toxicological and epidemiological studies have linked PFAS exposure to adverse health effects. For example, the presence of PFOA and PFOS in human blood samples collected worldwide, at a concentration level of µg/L, has been demonstrated to be harmful, causing developmental and reproductive problems, immune system dysfunction, hormonal imbalances, etc. Due to their persistence and potential risks, PFASs are considered a major environmental and public health concern, making them a research focus for scientists, regulators, and policymakers all over the world. The U.S. EPA set a health advisory level of 0.070 μg/L for PFASs in 2016; this level was recently reduced to 0.004 ng/L for PFOA and 0.02 ng/L for PFOS [3]. The EU Water Framework Directive has proposed a limit value of 0.1 µg/L for twenty PFASs in total [4]. In many cases, however, the PFAS levels in drinking water sources surpass the safety threshold, necessitating treatment methodologies for PFASs in water to mitigate their adverse effects.

Recent years have witnessed a boom in studies of PFASs, including strategies for their control in water, as displayed in Figure 1b,c. These strategies can be categorized into three principal types, i.e., physical separation [5,6], biological treatment [7], and chemical degradation (oxidation and reduction processes) [8,9,10,11]. Physical removal technologies, including adsorption, ion exchange, reverse osmosis, and nanofiltration, either have low PFAS elimination efficiency or high cost [6]. These non-destructive processes may generate waste such as spent adsorbents that can give rise to re-contamination with their re-entry into the environment. Biological degradation and chemical degradation are destructive technologies that permanently remove PFASs from water, but biodegradations are incomplete, have slow rates, and are highly dependent on environmental conditions [12]. Traditional wet chemical oxidation can barely break the very strong C–F bonds, whereas advanced oxidation processes (AOPs), originally introduced by Glaze in 1987 and distinguished by employing free radicals as highly reactive oxidant species, are considered to be highly efficient and have strong potential for the complete mineralization of PFASs [13,14]. To overcome the high overpotential of defluorination from C–F bonds and accelerate decomposition, more effective AOPs have been developed by introducing catalysts and external energy (light, electricity, heat, ultrasound, etc.) to lower the activation energy and overpotential of PFAS degradation reactions [15,16,17,18]. Energy in the forms of light and electricity is more appropriate for in situ use, as such approaches are relatively low-cost and environmentally friendly. Therefore, photocatalytic and electrocatalytic AOPs in heterogeneous systems are promising for the efficient and complete removal of PFASs in practice. These AOPs have recently attracted considerable research interest, with a vast body of relevant literature [19,20,21,22,23,24,25,26,27,28].

Numerous review articles have provided broad overviews of different water treatment methods for PFAS remediation [29,30,31,32,33,34,35], among which some have reviewed general AOPs for PFAS degradation [34,35]. In contrast, the present review will focus on heterogeneous photo/electrocatalytic AOPs. This is a group of promising treatment methods with enhanced PFAS mineralization is characterized by the involvement of catalyst materials and optical/electrical energy. Since comprehensive analyses of the similarities and differences of these processes in terms of PFAS degradation are still rare, this review aims to give overall and targeted insights into the photocatalytic and electrocatalytic degradation of PFASs by heterogeneous AOPs, from the fundamental mechanisms to catalyst designs. This review may facilitate further research on developing advanced photocatalysts and electrocatalysts for efficient PFAS removal by AOPs.

2. Fundamentals of PFAS Degradation by AOPs

In comparison with conventional homogeneous AOPs, heterogeneous AOPs for PFAS degradation are improved systems with incorporated catalysts. Heterogeneous AOPs use catalysts to activate oxidating agents (e.g., H₂O₂, O₃, persulfate), which contribute to the generation of powerful reactive oxidant species (ROS) such as hydroxyl radicals (•OH) and sulfate radicals (SO₄⁻•) [36,37]. Defluorination and the elimination of head groups (e.g., carboxylate and sulfonate groups) are linked, depending on the dominant ROS, contributing to the loss of head groups or breaking C−F bonds in perfluoroalkyl chains [38]. Identifying effective ROS and their transformation pathways during oxidation reactions may provide insights into PFAS degradation mechanisms, which are fundamental for the design of heterogeneous AOPs [39].

2.1. General Pathways and ROS for PFAS Degradation

The degradation by AOPs of PFASs occurring in water environments starts with their hydrolysis products (Equation (1)). Below, perfluorocarboxylic acid (PFCA, C_nF_2n+1COOH) represents the PFAS, while the effective ROS is represented by •OH. The degradation mechanisms of PFASs by AOPs are closely linked to ROS and their effective sites of action. However, these pathways can be summarized as a general reaction course [40]. Specifically, the oxidation of C_nF_2n+1COO⁻ to C_nF_2n+1COO• (Equation (2)) is initiated by radicals (e.g., •OH) generated from activated oxidizing agents [41]. Light and electrical energy can also trigger this reaction, utilizing, for example, photogenerated holes (h⁺) under ultraviolet radiation or the electron transfer process at the anode [42]. Subsequently, spontaneous decarboxylation of C_nF_2n+1COO• occurs due to its instability (Equation (3)), and the resulting perfluoralkyl radicals C_nF_2n+1• transform into C_nF_2n+1OH via hydroxylation (Equation (4)) [43]. After the spontaneous elimination of HF from C_nF_2n+1OH (Equation (5)), the resulting acyl halide, C_n−1F_2n−1COF, undergoes a hydrolysis process (Equation (6)), generating a short-chain PFCA (C_n−1F_2n−1COO⁻). Afterwards, this decarboxylation-hydroxylation-elimination-hydroxylation (DHEH) procedure is performed repeatedly, with the remove of a CF₂ unit and the release of CO₂ and HF in each cycle (Equations (2)–(6)) until complete mineralization is achieved [44,45]. The general degradation pathways of perfluorosulfonic acid (PFSA) are similar, but with the initial oxidation of C_nF_2n+1SO₃⁻ to C_nF_2n+1• by ROS attacking the C−S bond [46]. The resulting product then enters the defluorination cycle, similarly to PFCA.

$$C_n{F}_{2n+1}COOH\to C_n{F}_{2n+1}COO{}^{-}+H^{+}$$

(1)

$$C_n{F}_{2n+1}CO{O}^{-}+•OH(/+h^{+}/-e^{-})\to C_n{F}_{2n+1}COO•+O{H}^{-}$$

(2)

$$C_n{F}_{2n+1}COO•\to C_n{F}_{2n+1}•+C{O}_2$$

(3)

$$C_n{F}_{2n+1}•+•OH\to C_n{F}_{2n+1}OH$$

(4)

$$C_n{F}_{2n+1}OH\to C_{n-1}{F}_{2n-1}COF+HF$$

(5)

$$C_{n-1}{F}_{2n-1}COF+H_{2}O\to C_{n-1}{F}_{2n-1}CO{O}^{-}+HF+H^{+}$$

(6)

Effective ROS in AOPs of PFASs vary based on the oxidizing agents and their activation methods. Free radicals, such as •OH, sulfate radicals (SO₄⁻•), and superoxide radicals (•O₂⁻), as well as nonradicals like singlet oxygen (¹O₂) and holes (h⁺), have been identified as dominant ROS that contribute individually or synergistically to the defluorination and destruction of PFASs through advanced oxidation [47]. Fenton and Fenton-like processes, using hydrogen peroxide (H₂O₂) as the oxidant, generate •OH as the effective ROS when activated by Fe²⁺ or other transition metal-based chemicals or materials. Activated persulfate systems, utilizing persulfate (PDS, S₂O₈²⁻) or peroxymonosulfate (PMS, HSO₅⁻), primarily generate SO₄⁻• as the dominant ROS and exhibit reactivity due to the high redox potential (+ 2.5 V~3.1 V) and long lifetime (3.4 × 10⁻⁵ s) of SO₄⁻• [48,49]. Meanwhile, it should be noted that the activation methods of PMS make a big difference to the types of ROS, e.g., SO₄⁻• via activation with carbon materials, SO₄⁻• and •OH via thermal and radiation activation, SO₄⁻• and peroxymonosulfate anion radicals (SO₅•⁻) via transition metal activation, and •O₂⁻ and ¹O₂ via alkali activation [50]. In addition, photogenerated h⁺ acts as a powerful ROS that directly oxidizes PFASs or converts H₂O/O₂ to •OH/•O₂⁻, thereby generating more ROS [28]. Electrochemical processes induce the rapid generation of different ROS at the anode or facilitate electron transfer to the anode [51]. Hence, the application of light or electrical energy in AOPs has the potential to promote PFAS degradation by providing more powerful ROS or accelerating the oxidation processes. Photocatalytic or electrocatalytic AOPs are promising systems for highly effective PFAS remediation, deserving in-depth analysis and further research.

2.2. Principles of Photocatalytic AOPs for PFAS

Photocatalytic AOPs for FPAS are developed based on direct photo-degradation, achieved by breaking apart the C−F bonds using light of a specific wavelength. The direct photolysis process requires a match between the adsorption spectrum of the chemical bonds and the emission spectrum of the light, with the wavelength of the light playing a crucial role [52]. For example, PFOA has demonstrated strong UV adsorption and fast degradation at 185 nm [53,54], while light with a wavelength above 220 nm is barely absorbed by PFASs [29,55]. In photocatalytic AOPs of PFASs, the indirect photo-oxidation process is characterized by decarboxylation followed by defluorination, believed to be related to the photoinduced holes that exhibit a strong oxidizing capacity for organics. These holes work synergistically with other ROS to enhance PFAS degradation [56,57]. The system of photocatalytic AOPs consists of three components: the light, oxidant, and photocatalyst. There are two principles of PFAS degradation in photocatalytic AOPs: direct oxidation by photogenerated holes and co-oxidation with other ROS that are generated at the surface of catalysts with the assistance of the holes. The general process of PFAS degradation in photocatalytic AOPs can be described as follows: (1) Catalysts absorb light with energy (hv) equal to or greater than the band gap, which excites electrons from the valence band (VB) to the conduction band (CB), creating holes in the VB. (2) The generated electron−hole (e⁻−h⁺) pairs migrate to the surface of catalysts and react with the adsorbed PFAS. (3) The e⁻−h⁺ pairs react with precursors and generate ROS which assist in PFAS decomposition. For example, h⁺ and e⁻ react with H₂O and O₂, respectively, to produce •OH and •O₂⁻ (Equations (7) and (8)) [58,59].

$$h^{+}+H_{2}O\to •OH+H^{+}$$

(7)

$$e^{-}+O_2\to •O_2{}^{-}$$

(8)

The general mechanism of photocatalytic AOPs for PFAS degradation is summarized in Figure 2, where the PFAS is represented by PFAC. Photocatalysts play a crucial role in this process, as they are responsible for generating effective ROS and binding PFAS molecules. Both of these factors determine the efficiency of degradation [46]. Therefore, the construction and structure engineering of photocatalysts have garnered significant research interest.

2.3. Principles of Electrocatalytic AOPs for PFASs

There are two types of electrocatalytic AOPs for PFASs: direct electro-oxidation and indirect electrochemical oxidation. Direct electro-oxidation is a simple AOP that occurs on the surface of an electrode (anode) with a direct transfer of electrons. It relies on the in situ generation of ROS (e.g., •OH) or the direct transfer of electrons from the PFAS to the anode [60]. On the other hand, indirect electrochemical oxidation processes are the primary electrocatalytic AOP for organics treatment. Unlike direct electro-oxidation, electrons in this process act as mediators or assist in the generation of powerful ROS [51]. For example, the degradation of PFASs starts with the release of electrons, forming C_nF_2n+1COO• (Equation (2)); this occurs under an anode potential higher than the oxidation potential of the PFAS [61]. Additionally, electrocatalytic AOPs can produce radicals and oxidants during the electrode process. This includes •OH, which is strongly adsorbed onto the anode surface (M), as shown in Equation (9), H₂O₂ from the dimerization of •OH (Equation (10)), and ozone (O₃) from the discharge of water molecules (Equation (11)) [62]. These products are highly reactive with certain intermediate products during the decarboxylation and defluorination processes of PFASs, contributing to the efficiency of degradation.

$$M+H_{2}O\to M(•OH)+H{}^{+}+e{}^{-}$$

(9)

$$2M(•OH)\to 2MO+H_2{O}_2$$

(10)

$$3H_{2}O\to O_3+6H^{+}+6e^{-}$$

(11)

The general mechanism of electrocatalytic AOPs for PFAS degradation is illustrated in Figure 3. According to previous studies, after the formation of C_nF_2n+1• [63], there several approaches for further mineralization. Process (a) and (b) initially undergo a reaction with •OH to form C_nF_2n+1OH (Equation (4)). In process (a), reactions from Equation (12) to Equation (13) occur [64], whereas in process (b), reactions from Equation (14) to Equation (16) take place [65,66]. Aside from •OH, other anodic ROS such as O₂ also react with C_nF_2n+1• (Equation (17)), and the oxidation product C_nF_2n+1OO• can react with other perfluoro-alkoxy radicals (R_FCOO•) (Equation (18)). The resulting C_nF_2n+1O• then decomposes into C_n−1F_2n−1• for further degradation in a new cycle (Equation (19)) [63,64,65,67,68]. The third pathway (c) follows the reactions from Equation (17) to Equation (19).

$$C_n{F}_{2n+1}OH+•OH\to C_n{F}_{2n+1}O•+H_{2}O$$

(12)

$$C_n{F}_{2n+1}O•\to C_{n-1}{F}_{2n-1}•+C{F}_{2}O$$

(13)

$$C_n{F}_{2n+1}OH\to C_{n-1}{F}_{2n-1}COF+HF$$

(14)

$$C_{n-1}{F}_{2n-1}COF+•OH\to C_n{F}_{2n}{O}_{2}H•$$

(15)

$$C_n{F}_{2n}{O}_{2}H•\to C_{n-1}{F}_{2n-1}COO•+HF$$

(16)

$$C_n{F}_{2n+1}•+O_2\to C_n{F}_{2n+1}OO•$$

(17)

$$C_n{F}_{2n+1}OO•+R_{F}COO•\to C_n{F}_{2n+1}O•+R_{F}CO•+O_2$$

(18)

$$C_n{F}_{2n+1}O•\to C_{n-1}{F}_{2n-1}•+C{F}_{2}O$$

(19)

In electrocatalytic processes, oxidation primarily occurs on the anode. Therefore, the choice of anode materials (i.e., electrocatalysts) plays a crucial role in electrocatalytic AOPs for PFAS degradation. The behavior of PFAS degradation can vary depending on the type of anode material used. Anode materials are classified into two types based on the interactions between the adsorbed •OH on the anode surface and the degradation of organics: active anodes and nonactive anodes. Active anodes, such as Ti/SnO₂−Sb/MnO₂ [63], have a low potential for O₂ evolution and are distinguished from non-active anodes, such as Ti/SnO₂−Sb [69], by their ability to transform M(•OH) into strong oxidants. Generally, anode materials with higher O₂-evolution potential exhibit weaker interactions between M(•OH) and their surface, but they have higher reactivity towards PFASs [51]. Therefore, the structures of anode materials are worth studying in detail to enhance the degradation of PFASs in electrocatalytic AOPs. Analyzing previous related works could contribute to further research in this area.

3. Photocatalysts in AOPs for PFAS Degradation

In photocatalytic AOPs for PFASs, catalysts play a crucial role in absorbing light energy and bandgap excitation, leading to the generation of effective ROS [46]. Based on the mechanisms of photocatalytic AOPs for PFAS degradation, strategies for constructing photocatalysts to enhance the degradation process can be summarized as follows: (1) increase the yield of ROS through element doping, introducing heterojunctions, etc. [26,70,71,72,73]; (2) increase the rate of reaction between ROS and PFASs by using composite materials and controlling their morphology [38]; and (3) expand degradation pathways by generating multiple ROS from advanced photocatalysts [74]. Currently, commonly used photocatalysts for PFAS degradation include titanium dioxide (TiO₂), indium oxide (In₂O₃), gallium oxide (Ga₂O₃), bismuth (Bi)-based materials, and their composites [27,75]. The structure engineering of these photocatalysts varies based on their intrinsic properties and elemental compositions. Therefore, the main catalysts in photocatalytic AOPs for PFAS degradation are analyzed in groups, comprising metal oxide-based materials, Bi-based materials, and other compounds and composites. Important studies on the photocatalysts used in the photocatalytic advanced oxidation of PFASs are summarized in Table 1.

3.1. Metal Oxide-Based Materials

Metal oxides, such as TiO₂, In₂O₃, and Ga₂O₃, have long been used as traditional semiconductors in the photocatalytic degradation of organics in water. These metal oxides have been extensively studied for PFAS degradation [27,46,76,77]. TiO₂-based materials, in particular, have been widely used as photocatalysts since the discovery of water splitting on a TiO₂ anode by Fujishima and Honda in 1972 [78]. Though TiO₂ has shown promise in heterogeneous photocatalysis due to its strong UV absorption, non-toxicity, and long-term photostability, it is not efficient for photocatalytic PFAS degradation. This is due to its narrow spectral range, wide bandgap (3.0 eV for the rutile phase and 3.0 eV for the anatase phase), low electron-hole separation efficiency, and poor adsorption performance. Therefore, modifications of TiO₂ are necessary to enhance its photocatalytic activity. Strategies for modification include metal/nonmetal element doping, carbon material loading, and heterostructure construction. To date, doping with Fe [79], Cu [79], Pb [24,76], Pt [80], Pd [81], Ag [82] in TiO₂ for enhanced PFAS degradation has been studied, as well as the co-doping of metals, such as Fe/Nb [70]. Metal doping involves controlling the doping amount and regulating the pH of the solution to avoid the competitive adsorption of OH⁻ on the catalyst surface under alkaline conditions, which may affect the PFAS treatment efficiency; see mechanism in Figure 4a. Doping non-metal elements into the crystal lattice of TiO₂ can significantly enhance its visible light activity with a reduced bandgap width by changing the positions of the CB and VB, thereby improving the degradation efficiency of PFASs in photocatalysis [83]. Carbon materials, such as carbon nanotubes [84,85], graphene-based materials [86,87,88,89], and activated carbon [90], have also been employed to enhance the photocatalytic degradation of PFASs when loaded onto modified TiO₂. These carbon materials allow for uniform dispersion of TiO₂ nanoparticles on a hydrophobic surface, leading to an increased PFAS adsorption rate and improved photocatalytic degradation with minimized risk of secondary pollution and good stability. Heterojunctions between TiO₂ and other semiconductors can enhance the electron-hole separation efficiency and improve the photocatalytic degradation performance [26,28,71]. For instance, Sb₂O₃−TiO₂ heterojunctions use Sb₂O₃ nanoparticles confined within the mesoporous TiO₂ framework to adjust the band structure, increase the number of active sites for PFAS degradation, enhance UV absorption, and improve light utilization (Figure 4b) [71]. BN/TiO₂ heterojunctions facilitate charge carrier separation and enhance the degradation rate of PFOA compared to TiO₂ alone (15 times faster) [28].

In₂O₃ is a PFAS affinity material with a narrow bandgap of 2.8 eV, exceptional photocatalytic activity, and sensitivity to visible light. When compared to TiO₂, In₂O₃ has shown a remarkable 8.4-fold increase in the degradation rate coefficient of a PFAS (PFOA) under UV irradiation. These findings suggest it is a promising photocatalyst for PFAS decomposition [46]. Modifications are necessary for In₂O₃ due to its limitations, i.e., its low specific surface area and the rapid recombination of photogenerated electron-hole pairs. One effective approach is the generation of oxygen vacancies on the In₂O₃ surface, which enhances its photocatalytic performance. Additionally, nanostructures like nanospheres [42], (porous) nanosheets, and nanocubes [91,92] have been developed to provide adsorption sites for PFOA and oxygen atom binding sites in the carboxyl groups. These modifications ultimately contribute to the improved photocatalytic decomposition of PFOA. Several composite materials have been reported, such as g-C₃N₄−In₂O₃ [93], CeO₂−In₂O₃ [94], and MnOx−In₂O₃ [74]. Among them, the deposition of MnOx onto In₂O₃ surfaces has shown great potential; it creates abundant surface oxygen vacancies in In₂O₃, leading to the generation of active species and enhanced absorption of solar light (Figure 4c). This trend reflects the ongoing research efforts to develop photocatalytic materials for the enhanced degradation of PFASs.

Ga₂O₃ has excellent conductivity and tunable optical properties, despite its wide band-gap (4.9 eV). Studies have demonstrated its remarkable UV photocatalytic activity against PFASs, specifically in the context of PMS-assisted photocatalytic AOPs. The Ga₂O₃/PMS/UV system, with SO₄⁻• and •O₂⁻ as key ROS, achieves 100% degradation within 60 min [38]. Current research on Ga₂O₃ focuses on structure engineering to enhance its photocatalytic activity, primarily through size and morphology control. For example, synthesizing the compound into nanoparticles, monoclinic rod-like crystals, needle-like structures, and sheet-like structures has shown promising results. Furthermore, modifications have been explored to further improve the photocatalytic activity of Ga₂O₃. One common approach is metal-doping, such as Sn-doped β-Ga₂O₃ [72] and In-doped Ga₂O₃ (Figure 4d) [95]. Metal-doping on Ga₂O₃ promotes photocatalytic degradation by enhancing absorption through the hole oxidation process. This strategy provides a new method for removing PFOA from different water sources, capitalizing on the strong bonding ability between metal-doped Ga₂O₃ and PFASs.

3.2. Bi-Based Materials

Bismuth (Bi)-based photocatalysts have emerged as one of the most promising photocatalytic materials for catalysis, primarily due to their non-toxicity, high stability, and low cost. Commonly used Bi-based compounds for PFAS photocatalytic AOPs include bismuth oxyhalides (BiOX, where X is Cl, Br, I), bismuth ferrite (BiFeO₃, BFO), bismuth phosphate (BiPO₄), and bismuth hydroxyphosphate (Bi₃O(OH)(PO₄)₂, BiOHP); Table 1 summarizes their photocatalytic degradation efficiency with PFASs.

BiOX is a 2D layered compound with alternating double X ion layers and Bi₂O₂ layers along the c-axis. An internal electric field is formed between the halide planes and Bi₂O₂ layers, promoting faster charge transfer, enhanced redox potential, and excellent photocatalytic performance for effective PFAS degradation [39]. BiOI has a narrow bandgap (E_g = 1.67~1.92 eV) and high visible light absorption, showing great potential for applications. However, the narrow bandgap leads to the easy recombination of photoinduced e⁻/h+ pairs, which affects the photocatalytic activity [96]. Researchers have synthesized Br-doped BiOI (BiOI_0.95Br_0.05) for the photocatalytic degradation of PFOA [97]. Br doping not only increases PFOA adsorption but also expands the UV absorption range, leading to a significant enhancement in photocatalytic activity. However, the impact of non-metals on the structural properties of the semiconductor, in terms of stability, reproducibility, loading capacity, and practical applications of BiOI, requires further investigation. BiOCl, another noteworthy semiconductor with an indirect bandgap (E_g ranging from 2.62 to 3.46 eV) possesses excellent electronic and optical properties and has become a popular photocatalytic material [98]. Researchers have successfully prepared oxygen-deficient BiOCl nanosheets employing a simple hydrolysis method [99] and a rapid microwave-assisted solvothermal method [100]. These synthesis routes enable tight carboxyl group binding at the end of PFOA through monodentate and bidentate coordination, resulting in a significant improvement in defluorination effectiveness. Notably, defluorination rates achieved through a microwave-assisted solvothermal method surpasses the conventional solvothermal method by 2.7 and 33.8 times. In addition, the incorporation of ZnAl−LDHs with BiOCl (B−BHZA) has proven to be a promising strategy, as it lowers the bandgap energy of BiOCl and introduces abundant surface defects, facilitating charge generation and separation at the heterojunction [101]. Moreover, B−BHZA extends the spectral range and enhances UV light absorption, thereby promoting the direct oxidation of PFASs by h⁺ species.

BiOHP exhibits superior catalytic activity towards PFASs under UVC ultraviolet radiation compared to BiPO₄. This enhanced performance can be attributed to the positively charged surface of hydroxylated BiOHP, which enables the adsorption of deprotonated PFOA without relying on photocarrier reactions. However, the specific mechanism underlying this phenomenon requires further investigation [102]. It should be noted that while BiOHP effectively degrades low concentrations of PFOA in groundwater, its degradation rate is slower due to the presence of carbonate and natural organic matter [102]. To overcome this limitation, a combination of BiOHP with carbon spheres has been found to enhance its stability and catalytic performance. This synergistic effect is achieved through improved PFOA adsorption, facilitated electron transfer, and modified distribution of C−F bonds [103]. BFO is a perovskite-type mixed oxide that possesses a suitable bandgap for visible light degradation of organic pollutants. However, its photocatalytic performance is constrained, particularly for highly stable pollutants [104]. To enhance its catalytic activity, doping BFO with metals such as Pb has been investigated. The introduction of Pb provides reactive sites on the BFO surface, leading to improved performance [105]. In fact, combining Pb and reduced graphene oxide with BFO has been shown to enhance electron lifetime, facilitate reactive oxygen species generation, and promote the degradation of PFOA in water [106].

3.3. Other Compounds and Composites

In addition to metal-oxides and Bi-based materials, various photocatalysts have been employed in photocatalytic AOPs for PFAS degradation. These include metal/transition metal-based materials, metal-free materials, and modified composite materials. Qian et al. proposed a heterogeneous photocatalytic degradation mechanism for PFOA using Fe-zeolite under UVA irradiation (wavelength range: 320–420 nm) with O₂ as the terminal oxidant [107]. This Fe-zeolite catalyst, compared to homogeneous Fe³⁺, exhibits a broader light absorption range and can oxidize Fe²⁺ to Fe³⁺ in the presence of O₂, generating reactive species that contribute to PFAS mineralization. Other photocatalysts, such as samarium doped ferrite [108] and platinum-modified indium oxide nanorods (Pt/IONRs) [109], have also demonstrated significant PFOA degradation (48.6% and 98.0%) within a short period of time (1 h) in photocatalytic AOPs. The high degradation efficiency of Pt/IONRs can be attributed to Pt loading, the rod-like structure of the catalyst, and the presence of surface oxygen vacancies, which promote light harvesting, enhance the separation efficiency of the photogenerated charge carriers, and accelerate PFOA degradation. Furthermore, a synergistic effect of metal doping and carbon material loading has been observed in Ga/TNTs@AC composites, resulting in highly enhanced PFOA degradation (Figure 4e). Carbon-based materials play a crucial role in improving photocatalytic activity by enhancing conductivity (electron transfer), as explained by Zhao et al. (Figure 4f) [110]. A few reports have investigated the use of non-metal materials for the photocatalytic degradation of PFASs. Adsorption removal methods, e.g., using magnetic mesoporous carbon nitride and powdered activated carbon, have received significant attention. Interestingly, boron nitride (BN), which is a wide bandgap semiconductor with an energy gap of 6.0 eV, has shown promising heterogeneous photocatalytic activity toward PFOA [111]. BN, subjected to ball milling, exhibits a degradation rate four times higher than that of commercial TiO₂ under UV conditions. The photodegradation of PFOA by BN follows an oxidation mechanism involving h⁺, accompanied by degradation processes involving •O₂⁻ and •OH.

Figure 4. (a) The mechanism of metal-doping contributing to the photocatalytic degradation of PFOA on TiO₂ [80]. (b) The enhanced photocatalytic degradation of PFOA using a Sb₂O₃−TiO₂ heterojunction [71]. (c) MnO_x modification enhances the photocatalytic degradation of PFOA on In₂O₃ by introducing oxygen vacancies and generating more ROS [74]. (d) The band structures and photodegradation mechanism of PFOA on Ga₂O₃, with and without In-doping [95]. (e) The composite material Ga/TNTs@AC exhibits enhanced photodegradation of PFOSs through a synergistic effect of metal-doping and carbon material-loading [112]. (f) The enhanced photocatalytic degradation of PFOA has been achieved by loading a catalyst onto carbon aerogel (CA) over Ti [110].

Figure 4. (a) The mechanism of metal-doping contributing to the photocatalytic degradation of PFOA on TiO₂ [80]. (b) The enhanced photocatalytic degradation of PFOA using a Sb₂O₃−TiO₂ heterojunction [71]. (c) MnO_x modification enhances the photocatalytic degradation of PFOA on In₂O₃ by introducing oxygen vacancies and generating more ROS [74]. (d) The band structures and photodegradation mechanism of PFOA on Ga₂O₃, with and without In-doping [95]. (e) The composite material Ga/TNTs@AC exhibits enhanced photodegradation of PFOSs through a synergistic effect of metal-doping and carbon material-loading [112]. (f) The enhanced photocatalytic degradation of PFOA has been achieved by loading a catalyst onto carbon aerogel (CA) over Ti [110].

To conclude, various strategies in material construction and structure engineering have been developed to enhance the photodegradation of PFASs. These strategies aim to increase the adsorption of PFAS molecules onto catalysts or promote the generation of ROS. These approaches include:

(1)

The construction of defect sites through the introduction of oxygen vacancies and element doping. Defect sites serve as active sites for adsorption and catalytic reactions. Additionally, they broaden the light absorption range and enhance the light absorption capability by modifying the electronic and band structure.

(2)

The construction of heterojunctions by synthesizing composite materials. This involves combining two or more semiconductors with suitable band structures. The overlapping or coupling of these semiconductors enables the migration and separation of photo-generated charge carriers, effectively suppressing the recombination of electrons and holes. This leads to an enhanced photocatalytic efficiency.

(3)

The deposition of photocatalysts onto carbon material. The use of carbon nanomaterials, which possess a larger surface area and a porous structure, provides more active adsorption sites and improves catalyst stability. Furthermore, carbon materials facilitate interface charge transfer, prolong the lifetime of photo-generated charged carriers, and enhance visible light absorption by adjusting the band gap and acting as sensitizers.

(4)

The modulation of the crystal plane structure and electronic structure through the adjustment of the catalyst preparation methods. Exposing crystal planes with high surface energy and reactivity facilitates the separation of photo-generated electrons and holes, thus enhancing the photocatalytic activity of the catalysts.

Table 1. Summary of photocatalysts in photocatalytic AOPs for PFAS degradation.

Photocatalysts		Target PFAS	[PFAS]0	Experimental Conditions [Catalyst], Light Wavelength/Strength, pH	Removal Efficiency	Structural Engineering	Ref.
Metal oxide-based materials	TiO2	PFOA	59 mg·L−1	0.25 g·L−1, 400~770 nm/300 W, pH = 3 TiO2/UV/PMS system, ([PMS] = 0.75 g·L−1)	100% (8 h)	none	[75]
	TiO2 nanotubes	PFOA	50 mg·L−1	0.125 g·L−1, 254 nm/400 W, pH = 4	85% (24 h)	morphology control	[113]
	Pb−TiO2	PFOA	50 mg·L−1	0.5 g·L−1, 254 nm/400 W, pH = 5	50% (1.3 h)	metal-doping	[76]
	Pt−TiO2	PFOA	60 mg·L−1	0.5 g·L−1, 365 nm/125 W, pH = 3	100% (7 h)	metal-doping	[80]
	Cu−TiO2	PFOA	50 mg·L−1	0.5 g·L−1, 254 nm/100 W, pH = 5	91% (12 h)	metal-doping	[79]
	Sb2O3/TiO2	PFOA	10 mg·L−1	2.5 g·L−1, 200~280 nm/4 W, pH = 4.4	81.7% (2 h)	heterojunction	[71]
	BN/TiO2	PFOA	100 μM	0.5 g·L−1, 254 nm/4 W, pH = 3.2	97.6% (4 h)	heterojunction	[28]
	PLA−3DP TiO2	11PFASs	ng·L−1~mg·L−1	N/A, 280~400 nm/1.0 mW, pH = 7.1 ± 1	80% (24 h)	morphology control	[25]
	TiO2−MWCNTs	PFOA	30 mg·L−1	1.6 g·L−1, 365 nm/300 W, pH = 3	94% (8 h)	carbon material loading	[84]
	MWCNTs/C−TiO2	PFOA	2 mg·L−1	1.0 g·L−1, 420 nm/300 W, pH = 4.65	90% (3.5 h)	morphology control	[85]
	rGO/TiO2	PFOA	100 mg·L−1	1.0 g·L−1, 200~600 nm/150 W, pH = 7	93 ± 7% (12 h)	morphology control	[89]
	Ti3C2/TiO2	PFOA	20 μM	0.2 g·L−1, 254 nm/4.5 W, pH = 3	>99.9% (16 h)	heterojunction	[114]
	In2O3	PFOA	100 μM	0.5 g·L−1, 254 nm, pH = 4.2	80% (4 h)	none	[46]
	g-C3N4/In2O3	PFOA	200 mg·L−1	0.4 g·L−1, 254 nm/500 W, pH = N/A	91% (1 h)	carbon material loading	[93]
	In2O3−GRs	PFOA	30 mg·L−1	0.4 g·L−1, 254 nm/15 W, pH = N/A	100% (3 h)	carbon material loading	[115]
	CeO2/In2O3	PFOA	100 mg·L−1	0.4 g·L−1, 254nm/500 W, pH = 2.84	100% (1 h)	heterojunction	[94]
	MnOx−In2O3	PFOA	50 mg·L−1	0.5 g·L−1, visual light/500 W, pH = 3.8	99.8% (3 h)	heterojunction	[74]
	β-Ga2O3	PFOA	10 mg·L−1	0.5 g·L−1, 254 nm/50 W, pH = 7	98.8% (1.5 h)	none	[77]
	In−Ga2O3	PFOA	20 mg·L−1	0.5 g·L−1, 320 nm/200 W, pH = 7	100% (1 h)	metal-doping	[95]
	ZnO	PFOA	10 mg·L−1	0.2 g·L−1, 254 nm/28 W, pH = 4.5 ZnO/UV/O3 system	70.5% (4 h)	none	[116]
	ZnO−rGO	PFOA	10 mg·L−1	0.2 g·L−1, 254 nm, pH = N/A O3/UV/ZnO−rGO/S2O82− system	99.2% (4 h)	carbon material loading	[117]
Bi-based materials	BiOX/TiO2	PFOA	10 mg·L−1	0.2 g·L−1, 254 nm/30 W, pH = 7	100% (8 h)	carbon material loading	[26]
	BiOCl nanosheets	PFOA	0.02 mM	0.5 g·L−1, 254 nm/10 W, pH = 4.8	59.3% (24 h)	morphology control	[99]
	BiOI0.95Br0.05	PFOA	20 mg·L−1	0.4 g·L−1, 254 nm/300 W, pH = 7	96% (2 h)	crystal facet control	[97]
	BiOI/Bi5O7I	PFOA	15 mg·L−1	0.5 g·L−1, 400~760 nm/800 W, pH = 3.0	80% (6 h)	heterojunction	[73]
	BiOCl/BiPO4	PFOA	20 mg·L−1	0.5 g·L−1, 254 nm/2 W, pH = 7	100% (45 h)	heterojunction	[118]
	BiOHP	PFOA	0.5 mg·L−1	1.8 g·L−1, 254 nm/18 W, pH = 4.0	70%(20 min)	morphology control	[102]
	BiOHP/CS	PFOA	0.2 mg·L−1	1.0 g·L−1, 254 nm/18 W, pH = 7.0	>90% (1 h)	carbon material loading	[103]
Other composites	Fe−BEA35	PFOA	48 μM	1 g·L−1, 365 nm/4 W, 254 nm/4 W, pH = 3	>99% (24 h)	metal-doping + morphology control	[107]
	Pt/IONRs	PFOA	200 mg·L−1	0.4 g·L−1, 254 nm/500 W, pH = 1.85	98% (1 h)	metal-doping + morphology control	[119]
	Fe/TNTs@AC	PFOA	0.1 mg·L−1	1.0 g·L−1, 254 nm/21 W, pH = 3.0	90% (4 h)	metal-doping + carbon material loading	[90]
	Ga/TNTs@AC	PFOS	100 μg·L−1	1 g·L−1, nm/200 W, pH = 7.0 ± 0.1	75% (4 h)	metal-doping + carbon material loading	[112]
	Zn−AlLDHs −BiOCl	PFOA	0.5 mg·L−1	0.5 g·L−1, <350 nm/50 W, pH = 2.0	90% (6 h)	metal-doping + morphology control	[101]
	Pb−BiFeO3 /rGO	PFOA	50 mg·L−1	0.1·L−1, 254 nm/5 W, pH = 2.0	69.6% (8 h)	metal-doping + carbon material loading	[106]

Table 1. Summary of photocatalysts in photocatalytic AOPs for PFAS degradation.

Photocatalysts		Target PFAS	[PFAS]0	Experimental Conditions [Catalyst], Light Wavelength/Strength, pH	Removal Efficiency	Structural Engineering	Ref.
Metal oxide-based materials	TiO2	PFOA	59 mg·L−1	0.25 g·L−1, 400~770 nm/300 W, pH = 3 TiO2/UV/PMS system, ([PMS] = 0.75 g·L−1)	100% (8 h)	none	[75]
	TiO2 nanotubes	PFOA	50 mg·L−1	0.125 g·L−1, 254 nm/400 W, pH = 4	85% (24 h)	morphology control	[113]
	Pb−TiO2	PFOA	50 mg·L−1	0.5 g·L−1, 254 nm/400 W, pH = 5	50% (1.3 h)	metal-doping	[76]
	Pt−TiO2	PFOA	60 mg·L−1	0.5 g·L−1, 365 nm/125 W, pH = 3	100% (7 h)	metal-doping	[80]
	Cu−TiO2	PFOA	50 mg·L−1	0.5 g·L−1, 254 nm/100 W, pH = 5	91% (12 h)	metal-doping	[79]
	Sb2O3/TiO2	PFOA	10 mg·L−1	2.5 g·L−1, 200~280 nm/4 W, pH = 4.4	81.7% (2 h)	heterojunction	[71]
	BN/TiO2	PFOA	100 μM	0.5 g·L−1, 254 nm/4 W, pH = 3.2	97.6% (4 h)	heterojunction	[28]
	PLA−3DP TiO2	11PFASs	ng·L−1~mg·L−1	N/A, 280~400 nm/1.0 mW, pH = 7.1 ± 1	80% (24 h)	morphology control	[25]
	TiO2−MWCNTs	PFOA	30 mg·L−1	1.6 g·L−1, 365 nm/300 W, pH = 3	94% (8 h)	carbon material loading	[84]
	MWCNTs/C−TiO2	PFOA	2 mg·L−1	1.0 g·L−1, 420 nm/300 W, pH = 4.65	90% (3.5 h)	morphology control	[85]
	rGO/TiO2	PFOA	100 mg·L−1	1.0 g·L−1, 200~600 nm/150 W, pH = 7	93 ± 7% (12 h)	morphology control	[89]
	Ti3C2/TiO2	PFOA	20 μM	0.2 g·L−1, 254 nm/4.5 W, pH = 3	>99.9% (16 h)	heterojunction	[114]
	In2O3	PFOA	100 μM	0.5 g·L−1, 254 nm, pH = 4.2	80% (4 h)	none	[46]
	g-C3N4/In2O3	PFOA	200 mg·L−1	0.4 g·L−1, 254 nm/500 W, pH = N/A	91% (1 h)	carbon material loading	[93]
	In2O3−GRs	PFOA	30 mg·L−1	0.4 g·L−1, 254 nm/15 W, pH = N/A	100% (3 h)	carbon material loading	[115]
	CeO2/In2O3	PFOA	100 mg·L−1	0.4 g·L−1, 254nm/500 W, pH = 2.84	100% (1 h)	heterojunction	[94]
	MnOx−In2O3	PFOA	50 mg·L−1	0.5 g·L−1, visual light/500 W, pH = 3.8	99.8% (3 h)	heterojunction	[74]
	β-Ga2O3	PFOA	10 mg·L−1	0.5 g·L−1, 254 nm/50 W, pH = 7	98.8% (1.5 h)	none	[77]
	In−Ga2O3	PFOA	20 mg·L−1	0.5 g·L−1, 320 nm/200 W, pH = 7	100% (1 h)	metal-doping	[95]
	ZnO	PFOA	10 mg·L−1	0.2 g·L−1, 254 nm/28 W, pH = 4.5 ZnO/UV/O3 system	70.5% (4 h)	none	[116]
	ZnO−rGO	PFOA	10 mg·L−1	0.2 g·L−1, 254 nm, pH = N/A O3/UV/ZnO−rGO/S2O82− system	99.2% (4 h)	carbon material loading	[117]
Bi-based materials	BiOX/TiO2	PFOA	10 mg·L−1	0.2 g·L−1, 254 nm/30 W, pH = 7	100% (8 h)	carbon material loading	[26]
	BiOCl nanosheets	PFOA	0.02 mM	0.5 g·L−1, 254 nm/10 W, pH = 4.8	59.3% (24 h)	morphology control	[99]
	BiOI0.95Br0.05	PFOA	20 mg·L−1	0.4 g·L−1, 254 nm/300 W, pH = 7	96% (2 h)	crystal facet control	[97]
	BiOI/Bi5O7I	PFOA	15 mg·L−1	0.5 g·L−1, 400~760 nm/800 W, pH = 3.0	80% (6 h)	heterojunction	[73]
	BiOCl/BiPO4	PFOA	20 mg·L−1	0.5 g·L−1, 254 nm/2 W, pH = 7	100% (45 h)	heterojunction	[118]
	BiOHP	PFOA	0.5 mg·L−1	1.8 g·L−1, 254 nm/18 W, pH = 4.0	70%(20 min)	morphology control	[102]
	BiOHP/CS	PFOA	0.2 mg·L−1	1.0 g·L−1, 254 nm/18 W, pH = 7.0	>90% (1 h)	carbon material loading	[103]
Other composites	Fe−BEA35	PFOA	48 μM	1 g·L−1, 365 nm/4 W, 254 nm/4 W, pH = 3	>99% (24 h)	metal-doping + morphology control	[107]
	Pt/IONRs	PFOA	200 mg·L−1	0.4 g·L−1, 254 nm/500 W, pH = 1.85	98% (1 h)	metal-doping + morphology control	[119]
	Fe/TNTs@AC	PFOA	0.1 mg·L−1	1.0 g·L−1, 254 nm/21 W, pH = 3.0	90% (4 h)	metal-doping + carbon material loading	[90]
	Ga/TNTs@AC	PFOS	100 μg·L−1	1 g·L−1, nm/200 W, pH = 7.0 ± 0.1	75% (4 h)	metal-doping + carbon material loading	[112]
	Zn−AlLDHs −BiOCl	PFOA	0.5 mg·L−1	0.5 g·L−1, <350 nm/50 W, pH = 2.0	90% (6 h)	metal-doping + morphology control	[101]
	Pb−BiFeO3 /rGO	PFOA	50 mg·L−1	0.1·L−1, 254 nm/5 W, pH = 2.0	69.6% (8 h)	metal-doping + carbon material loading	[106]

4. Electrocatalysts in AOPs for PFAS Degradation

Electrocatalysts assume the role of electrode materials in electrocatalytic AOPs, where electrochemical oxidations for PFAS degradation occur on their surface. The key factors determining the efficiency of anodic oxidation in electrocatalytic AOPs include the production of ROS, the rate of mass transfer from the bulk solution to the electrode surface, the availability of active sites on the electrode surface for degradation reactions of PFASs and their transformations, as well as the stability and current transfer efficiency of the anode materials [22,120,121,122]. Based on these factors, the electrocatalytic performance of anode materials can be improved by increasing the number of active sites, enhancing the electron transfer rate, accelerating mass transfer, and employing strategies such as introducing defects, elemental doping, surface modifications, designing the microstructure of the electrode, and exposing preferred crystal facets [71,123,124]. Currently, a wide variety of photocatalysts are used for PFAS degradation, with commonly employed anode materials including boron-doped diamond (BDD), tin oxide (SnO₂), lead oxide (PbO₂), and others [63,66,125]. To clarify the strategies used in constructing materials for different groups of electrocatalysts, a classified analysis of recently studied electrocatalytic materials for PFAS degradation on the anode is provided, specifically focusing on BBD-based electrodes, metal oxide-based materials, and other hybrids materials. Important research on catalysts for the electrocatalytic degradation of PFASs is summarized in Table 2.

4.1. BDD-Based Materials

Boron-doped diamond (BDD) is a one of the most frequently used anode materials in the electrocatalytic degradation of PFASs, owing to its wide operational potential window, excellent chemical stability, and high oxidation potential (2.7 V vs. SHE) [65,126,127]. The BDD electrode plays a crucial role in the direct electrochemical oxidation of PFASs at low current densities, rather than relying on the •OH oxidation process at high current densities [11,128]. BDD films have been proven to be effective anodic electrocatalysts for PFOS degradation [127], with electron transfer from PFOS to the anode resulting in the formation of final products such as SO₄²⁻, F⁻, CO₂ and a small amount of trifluoroacetic acid. However, the widespread application of BDD as an anode material in electrocatalytic AOPs for PFASs is limited due to its high cost and a lack of suitable electrode substrates for BDD. Metals and nonmetals like Ta, Nb, W, and Si have been investigated as BDD substrates, and Si/BDD has been found to be a cost-effective option for PFOS degradation (>90%) [123,129]. Moreover, B/N co-doped diamond (BND) has been developed as an anode for sulfate-activated electrocatalytic AOPs for PFASs [120]. B/N co-doping enhances ROS generation by increasing the number of active sites, thereby promoting the electrochemical activation of the sulfate solution and PFAS degradation, as depicted in Figure 5a. However, the electrocatalytic activity of BDD is hindered by the occurrence of fluorination on the surface through the binding of fluoride ions in the solution; to date, only a handful of reports have addressed this issue [126,130]. Therefore, further research should be conducted on the surface engineering of BDD to reduce F⁻ adsorption on the anode surface and thereby enhance its efficiency in electrocatalytic PFAS degradation.

4.2. Metal Oxide-Based Materials

The utilization of metal oxides as electrode materials has revealed the drawbacks associated with metal electrodes, such as lower oxygen evolution overpotential and susceptibility to oxidation. Additionally, it has made it possible to overcome the limitations of BDD electrodes, including low conductivity and low efficiency in terms of utilizing •OH radicals [131]. Currently, metal oxides commonly used as electrode materials include SnO₂, PbO₂, TiO₂, MnO₂, and others. Among these, SnO₂ and PbO₂ demonstrate outstanding electrocatalytic oxidation performance, with a higher overpotential of O₂ evolution, making them suitable for the degradation of organics.

SnO₂, a semiconductor with a bandgap of 3.5eV, faces challenges in its direct use as an electrode material due to its high resistance. However, its conductivity can be improved through doping [22,122,132], particularly with antimony (Sb) [63,125]. In addition, SnO₂−Sb electrodes exhibit a high oxygen evolution potential, which contributes to their extensive application in PFAS degradation [63,65,110]. By doping Ti/SnO₂−Sb with multiple metals, a PFOA removal rate of 93.3% can be achieved [63,65]. It is important to note that the presence of SO₄²⁻ in the solution can cover the active sites on the electrode. This reduces the production of •OH radicals, subsequently decreasing the PFOA removal rate [110]. Moreover, F-doped SnO₂ electrodes (Ti/SnO₂−F) demonstrate a higher oxygen evolution potential and better catalytic activity than Ti/SnO₂−Sb electrodes. This can be attributed to the smooth surface of F-doped SnO₂ electrodes, which minimizes electrolyte infiltration and ensures electrode stability [122]. Furthermore, a Ti/SnO₂−Bi electrode achieved a 99% PFOA removal rate within 2 h [65]. However, the high anode potential (3.37 V) led to the primary degradation mechanism being the direct oxidation of PFOA on the anode through decarboxylation, resulting in inferior performance for PFOS degradation. Apart from Ti, carbon-based materials hold promise as substrates for metal oxide-based electrocatalysts. For example, carbon aerogel (CA)/SnO₂−Sb electrodes exhibit a significantly improved degradation rate for PFOA, i.e., 3.5 times higher than for Ti/SnO₂−Sb electrodes [110].

PbO₂ anodes offer several advantages for efficient PFAS degradation, such as low processing costs, simple preparation, high conductivity, and a high oxygen evolution potential [133,134]. However, the issue of PbO₂ detachment in Ti/PbO₂ electrodes hampers their stability. To address this, Ti/SnO₂−Sb/PbO₂, and TiO₂-Nanotubes (NTs)/Ag₂O/PbO₂ electrodes have been developed, effectively enhancing the stability and electrochemical degradation capability with the assistance of an interlayer [124,135,136]. A study was conducted [137], highlighting the contribution of interlayer metal/metal oxide anodes to the efficiency of PFOS degradation (Figure 5b). However, PbO₂ electrodes are susceptible to Pb²⁺ leaching [133,134]; this can be mitigated through doping with elements to reduce the grain size, increase the electroactive surface area, improve the oxygen evolution potential, and enhance electron migration ability. Doping with cerium (Ce), ytterbium (Yb), and zirconium (Zr) has resulted in removal rates of over 88% for PFOA [67,68,138]. A Ti/SnO₂−Sb/PbO₂−Ce electrode exhibited removal rates of over 92% for PFDA and PFNA. It is worth noting that the electrochemical degradation rate of PFASs on Ti/SnO₂-based electrodes is influenced by the chain length, emphasizing the need to tailor the catalyst according to the PFAS chain structure [63,67,139]. Carbon-based materials also contribute to electrocatalytic PFAS degradation as substrates for PbO₂. For instance, a 3D graphene (3DG)−PbO₂ anode obtained through electro-deposition exhibited a degradation rate constant for PFOS that was 2.33 times higher than that of PbO₂ anodes [140]. This can be attributed to the porous nanostructures, resulting in a larger specific surface area and multiple electronic transfer channels (Figure 5c) in the anode. Similarly, doping tetrafluoroethylene (PTFE) into PbO₂ enhances the mass transfer and •OH generation capacity. PTFE−PbO₂ provides more active sites and a faster electron transfer rate, contributing to outstanding degradation efficiency [22].

Other metal oxide-based electrocatalysts, such as Ti₄O₇, have recently been studied as promising electrocatalysts. The use of a porous Magnéli phase Ti₄O₇ membrane as an anode has been shown to improve the electro-degradation efficiency for PFOS [141]. This improvement can be attributed to the increased electroactive surface area and enhanced interphase mass transfer facilitated by the Ti₄O₇ reactive electrochemical membrane (REM). In addition, doping Ti₄O₇ with metal ions (e.g., Ce³⁺) and metals (e.g., amorphous Pd) has proven to be effective in enhancing the electrocatalytic degradation. Ce³⁺ ions promote ROS generation and electron transfer (Figure 5d) [121], while the high oxidation state and electron-deficient 5d orbitals of amorphous Pd clusters facilitate the efficient extraction of electrons from PFOA [142]. The abundant Pd−O species on the surface serve as channels for the transfer of electrons from PFOA to the electrode, thereby enhancing the anodic oxidation capability. Furthermore, Wang et al. developed Ti³⁺/TiO₂−nanotube arrays (NTA) with both metal ion doping and a 3D structure construction. This approach achieved enhanced electron and mass transfer, resulting in efficient electrocatalytic degradation for PFOA (Figure 5e) [21].

4.3. Other Compounds and Composites

Composites and hybrids used as anode materials for the electrocatalytic degradation of PFASs can take various forms, employing different mechanisms to improve electrocatalysis performance. However, most of the composites or hybrids used as electrocatalysts are either BDD-based or metal oxide-based materials, as discussed above. Furthermore, there are only a few studies available on this topic, leaving ample room for further research [19,20,143]. In terms of electrode materials, one scarcely investigated type which has demonstrated high reactivity and chemical robustness is the multifunctional single-atom catalyst (SAC). SACs show promise in electrocatalytic PFAS degradation [144]. In a recent study, a Co-based SAC (Co−CN₂) immobilized with Fe₂O₃ was reported as a highly efficient electrocatalyst for PFOA degradation. This was attributed to the construction of catalytic Fenton reactions with locally generated H₂O₂ (Figure 5f) [20]. In addition to anode materials, research on cathodes for enhanced electrocatalytic AOPs in PFAS degradation has also been conducted. These studies rely on the synergistic effect of cathodic electro-AOPs (Fenton) and anodic oxidation. For example, an Fe–Mn-based catalyst was developed by doping it into CA [143]. In this case, Fe and Mn were used as Fenton catalysts, providing better distribution of active sites for enhanced •OH generation. As a result, a 97% removal rate of PFOA was achieved with 4 h of electrocatalysis.

Figure 5. (a) The use of element (B/N)-doped diamond (BND) as an anode promotes PFOS electro-degradation by generating ROS [120]. (b) The Sn−Sb interlayer and F doping contribute to the degradation of PFOS on a Ti/SnO₂ anode [137]. (c) The 3DG-PbO₂ composite anode effectively degrades PFOSs due to its strong ROS generation capacity, abundant active sites, and small charge-transfer resistance [140]. (d) Ce³⁺ doping enhances the electrocatalytic degradation of PFOS on a Ti₄O₇ anode by increasing ROS generation and electron transfer [121]. (e) Enhanced electron and mass transfer on a Ti³⁺/TiO₂−NTA anode for PFOA mineralization is achieved through metal-doping and a 3D nanotube structure [21]. (f) The nearly complete mineralization of PFOA on Co−CN₂−Fe₂O₃ composites is attributed to the synergistic effect of single-atom catalysis on Co−CN₂ and Fenton catalysis on Fe₂O₃ [20].

Figure 5. (a) The use of element (B/N)-doped diamond (BND) as an anode promotes PFOS electro-degradation by generating ROS [120]. (b) The Sn−Sb interlayer and F doping contribute to the degradation of PFOS on a Ti/SnO₂ anode [137]. (c) The 3DG-PbO₂ composite anode effectively degrades PFOSs due to its strong ROS generation capacity, abundant active sites, and small charge-transfer resistance [140]. (d) Ce³⁺ doping enhances the electrocatalytic degradation of PFOS on a Ti₄O₇ anode by increasing ROS generation and electron transfer [121]. (e) Enhanced electron and mass transfer on a Ti³⁺/TiO₂−NTA anode for PFOA mineralization is achieved through metal-doping and a 3D nanotube structure [21]. (f) The nearly complete mineralization of PFOA on Co−CN₂−Fe₂O₃ composites is attributed to the synergistic effect of single-atom catalysis on Co−CN₂ and Fenton catalysis on Fe₂O₃ [20].

Based on a comprehensive analysis of key factors in the electrocatalytic oxidation of organics, including the generation of ROS, mass/charge transfer efficiency, and electrode stability, the structural engineering of electrocatalysts for effective PFAS degradation can be summarized as follows:

(1)

Enhancing the yield of ROS on the electrode by doping a specific functional substance onto the surface. Doping helps optimize the crystalline phase of the catalytic electrodes, promoting a more compact arrangement of particles and effectively enhancing the electrode’s ability to generate physically adsorbed ROS species.

(2)

Improving the surface mass transfer efficiency of the electrode through metal loading and the implementation of three-dimensional nanostructures. Metal-doping is an important strategy for reducing and improving electron transfer. Constructing 3D nanostructures on the electrode surface increases the specific surface area and active sites, thereby enhancing the migration of PFASs toward the electrode and improving their adsorption efficiency.

(3)

Enhancing the stability of the electrode by introducing functional intermediate layers. Incorporating the intermediate layer strengthens the interaction between the active layer and the substrate, preventing the detachment of the active layer, improving electrode stability, and extending the electrode’s lifespan.

Table 2. Summary of electrocatalysts in electrocatalytic AOPs for PFAS degradation.

Electrocatalysts		Target PFAS	[PFAS]0 (Electrolyte Solution)	Experimental Conditions Surface Area (S), Interelectrode Gap (L), Electric Current Density (J)	Removal Efficiency	Structural Engineering	Ref.
BDD-based materials	BDD	PFOS	0.4 mM (10 mM NaClO4)	S = 25 cm2, L = 2 mm, J = 20 mA·cm−2, T = 22 °C, pH = 4	50% (5.3 min)	none	[127]
	BDD (high boron)	PFOA +PFOS	0.1 mg L−1 (100 mM PBS)	S = 10.5 cm2, L = 2.5 cm, J = 75 mA·cm−2, T = 20~25 °C, pH = 7.8	80%PFOA, 78%PFOS	doping (B)	[11]
	BND	PFOA	50 mg L−1 (0.05 M Na2SO4)	S = 10.5 cm2, L = 2.5 cm, J = 4 mA·cm−2, pH = 4.8	77.4% (3 h)	doping (N)	[120]
	Si/BDD	PFOA	0.2 mg L−1 (0.4 g L−1 Na2SO4)	S = 81 cm2, L = 1 cm, J = 25 mA·cm−2, pH = 11	>90% (1 h)	metal-doping	[129]
	Ti/BDD	PFCAs	0.25 mM (10 mM NaClO4)	S = 25 cm2, L = 1.5 cm, J = 10 mA·cm−2, pH = 3	>95% (3 h)	metal-doping	[66]
Metal-oxide based materials	Ti3+/TiO2−NTA	PFOA	50 mg L−1 (20 mM Na2SO4)	S = 25 cm2, L = 1 cm, J = 2 mA·cm−2, pH = 3~11	98.1% (1.5 h)	doping + morphology control	[21]
	Ti/SnO2−Sb	PFOA	100 mg L−1 (10 mM NaClO4)	L = 1 cm, J = 10 mA·cm−2, T = 25 °C, pH = 5	90.3% (1.5 h)	co-doping	[63]
	Ti/SnO2−Sb/PbO2				91.1% (1.5 h)	metal-doping +intercalation
	Ti/SnO2−Sb/MnO2				31.7% (1.5 h)	metal-doping +intercalation
	Ti/SnO2−Sb/Bi2O3	PFOS	20 mg L−1 (1.4 g L−1 NaClO4)	S = 17.64 cm2, J = 6.8 mA·cm−2, T = 32 °C, pH = 6.94	23.8% (3.5 h)	metal-doping +intercalation	[145]
	Ti/SnO2−F	PFOA	100 mg L−1 (10 mM NaClO4)	S = 25 mm2, L = 1 cm, J = 100 mA·cm−2, T = 25 °C, pH = 7	96.5% (0.5 h)	co-doping	[122]
	SnO2−Sb/CA	PFOA	100 mg L−1 (0.1 M Na2SO4)	S = 5 cm2, L = 1 cm, J = 20 mA·cm−2, T = 20 °C, pH = 7	91% (5 h)	metal-doping +carbon material loading	[110]
	Ti/SnO2−Sb−Bi	PFOA	50 mg L−1 (1.4 g L−1 NaClO4)	S = 11.33 cm2, L = 10 mm, J = 22 mA·cm−2, pH = 4.71	>99% (2 h)	co-doping	[65]
	Ti/Sn−Sb/SnO2−F −Sb	PFOA	100 mg L−1 (10 mM NaClO4)	S = 25 mm2, L = 1 mm, J = 20 mA·cm−2, T = 25 ± 3 °C, pH = 2	99% (2 h)	co-doing +intercalation	[137]
	3DG−PbO2	PFOS	50 mg mL−1 (0.05 M Na2SO4)	J = 30 mA·cm−2, T = 30 °C, pH = 7	96.17% (2 h)	carbon material loading	[140]
	Ceramic/PbO2 −PTFE	PFOA	20 mg mL−1 (15 mM Na2SO4)	S = 20.6 cm2, L = 2 cm, J = 15 mA·cm−2, T = 25 °C, pH = 7	98.9% (5 h)	doping + loading	[22]
	Ti/TiO2NTs /Ag2O/PbO2	PFOS	0.0929 mM (1.4 g L−1 Na2SO4)	S = 12 cm2, L = 10 mm, J = 30 mA·cm−2, T = 30 ± 2 °C	74.87% (1.5 h)	metal-doping +morphology control	[146]
	Magnéli phase TinO2n–1	PFOS	2.0 μM (100 mM Na2SO4)	S = 4 cm2, L = 10 cm, J = 10 mA·cm−2, T = 25 ± 1 °C, pH = 6	98.30 ± 0.51% (2 h)	morphology control	[141]
	Porous Ti4O7	PFOA PFOS	0.5 mM PFOA, 0.1 mM PFOS (0.25 M Na2SO4)	S = 1 cm2, L = 1.5 cm, J = 10 mA·cm−2, T = 25 ± 1 °C,	99.9% PFOA 9.1% PFOS (3 h)	morphology control	[147]
	Ti4O7 REM	PFOA +PFOS	10 μM (100 mM K2HPO4)	S = 0.5 cm2, anode potential = 3.6 V/SHE, pH = 7	>99.9%	morphology control	[148]
	Ce-doped Ti4O7: (Ti1–xCex)4O7	PFOS	20 nM (10 mM Na2SO4)	S = 9 cm2, L = 5 mm, J = 20 mA·cm−2, pH = 7	>83.3% (2 h)	metal-doping	[121]
	Pd/Ti4O7	PFOA	0.12 mM (50 mM Na2SO4)	S = 25 cm2, V = 30 mL, J = 10 mA·cm−2, T = 25 ± 1 °C, pH = 7.2	>90% (1 h)	metal-doping	[142]
	Mixed metal oxide (MMO)	PFOA +PFOS	5 mg L−1 (500 mg/L Na2SO4)	S = 100 cm2, L = 16 mm, J = 10 mA·cm−2, pH = 7.4	>90% (8 h)	morphology control	[149]
	LaNixY1−xO3 (Y = Fe/Cu/Co/Sr)	PFOA	0.25 mM (0.05 M Na2SO4)	L = 3 cm, J = 20 mA·cm−2, pH = 5, 1.0 mM FeSO4	90% (Y = Sr, 2.5 h)	crystal facet control	[150]
Other composites	Mxene-based membrane	PFBA +PFOA	1 μg/L (0.1 M Na2SO4)	S = 12.56 cm2, L = 10 mm, J = 10 mA·cm−2, pH = 7.00 ± 0.10	>99% (3 h)	morphology control	[19]
	Co−CN2−Fe2O3	PFOA	1.9 mg L−1 (0.05 M Na2SO4)	S = 0.2475 cm2, J = 2.2 mA·cm−2, T = 25 °C, pH = 2	70% (30 min)	metal-doping +morphology control	[20]
	Fe−Mn/CA	PFOA	50 mg mL−1 (50 mM Na2SO4)	S = 7.0 cm2, L = 2 cm, J = 2.85 mA·cm−2, T = 25 °C, pH = 3	97% (4 h)	co-doping +morphology control	[143]

Table 2. Summary of electrocatalysts in electrocatalytic AOPs for PFAS degradation.

Electrocatalysts		Target PFAS	[PFAS]0 (Electrolyte Solution)	Experimental Conditions Surface Area (S), Interelectrode Gap (L), Electric Current Density (J)	Removal Efficiency	Structural Engineering	Ref.
BDD-based materials	BDD	PFOS	0.4 mM (10 mM NaClO4)	S = 25 cm2, L = 2 mm, J = 20 mA·cm−2, T = 22 °C, pH = 4	50% (5.3 min)	none	[127]
	BDD (high boron)	PFOA +PFOS	0.1 mg L−1 (100 mM PBS)	S = 10.5 cm2, L = 2.5 cm, J = 75 mA·cm−2, T = 20~25 °C, pH = 7.8	80%PFOA, 78%PFOS	doping (B)	[11]
	BND	PFOA	50 mg L−1 (0.05 M Na2SO4)	S = 10.5 cm2, L = 2.5 cm, J = 4 mA·cm−2, pH = 4.8	77.4% (3 h)	doping (N)	[120]
	Si/BDD	PFOA	0.2 mg L−1 (0.4 g L−1 Na2SO4)	S = 81 cm2, L = 1 cm, J = 25 mA·cm−2, pH = 11	>90% (1 h)	metal-doping	[129]
	Ti/BDD	PFCAs	0.25 mM (10 mM NaClO4)	S = 25 cm2, L = 1.5 cm, J = 10 mA·cm−2, pH = 3	>95% (3 h)	metal-doping	[66]
Metal-oxide based materials	Ti3+/TiO2−NTA	PFOA	50 mg L−1 (20 mM Na2SO4)	S = 25 cm2, L = 1 cm, J = 2 mA·cm−2, pH = 3~11	98.1% (1.5 h)	doping + morphology control	[21]
	Ti/SnO2−Sb	PFOA	100 mg L−1 (10 mM NaClO4)	L = 1 cm, J = 10 mA·cm−2, T = 25 °C, pH = 5	90.3% (1.5 h)	co-doping	[63]
	Ti/SnO2−Sb/PbO2				91.1% (1.5 h)	metal-doping +intercalation
	Ti/SnO2−Sb/MnO2				31.7% (1.5 h)	metal-doping +intercalation
	Ti/SnO2−Sb/Bi2O3	PFOS	20 mg L−1 (1.4 g L−1 NaClO4)	S = 17.64 cm2, J = 6.8 mA·cm−2, T = 32 °C, pH = 6.94	23.8% (3.5 h)	metal-doping +intercalation	[145]
	Ti/SnO2−F	PFOA	100 mg L−1 (10 mM NaClO4)	S = 25 mm2, L = 1 cm, J = 100 mA·cm−2, T = 25 °C, pH = 7	96.5% (0.5 h)	co-doping	[122]
	SnO2−Sb/CA	PFOA	100 mg L−1 (0.1 M Na2SO4)	S = 5 cm2, L = 1 cm, J = 20 mA·cm−2, T = 20 °C, pH = 7	91% (5 h)	metal-doping +carbon material loading	[110]
	Ti/SnO2−Sb−Bi	PFOA	50 mg L−1 (1.4 g L−1 NaClO4)	S = 11.33 cm2, L = 10 mm, J = 22 mA·cm−2, pH = 4.71	>99% (2 h)	co-doping	[65]
	Ti/Sn−Sb/SnO2−F −Sb	PFOA	100 mg L−1 (10 mM NaClO4)	S = 25 mm2, L = 1 mm, J = 20 mA·cm−2, T = 25 ± 3 °C, pH = 2	99% (2 h)	co-doing +intercalation	[137]
	3DG−PbO2	PFOS	50 mg mL−1 (0.05 M Na2SO4)	J = 30 mA·cm−2, T = 30 °C, pH = 7	96.17% (2 h)	carbon material loading	[140]
	Ceramic/PbO2 −PTFE	PFOA	20 mg mL−1 (15 mM Na2SO4)	S = 20.6 cm2, L = 2 cm, J = 15 mA·cm−2, T = 25 °C, pH = 7	98.9% (5 h)	doping + loading	[22]
	Ti/TiO2NTs /Ag2O/PbO2	PFOS	0.0929 mM (1.4 g L−1 Na2SO4)	S = 12 cm2, L = 10 mm, J = 30 mA·cm−2, T = 30 ± 2 °C	74.87% (1.5 h)	metal-doping +morphology control	[146]
	Magnéli phase TinO2n–1	PFOS	2.0 μM (100 mM Na2SO4)	S = 4 cm2, L = 10 cm, J = 10 mA·cm−2, T = 25 ± 1 °C, pH = 6	98.30 ± 0.51% (2 h)	morphology control	[141]
	Porous Ti4O7	PFOA PFOS	0.5 mM PFOA, 0.1 mM PFOS (0.25 M Na2SO4)	S = 1 cm2, L = 1.5 cm, J = 10 mA·cm−2, T = 25 ± 1 °C,	99.9% PFOA 9.1% PFOS (3 h)	morphology control	[147]
	Ti4O7 REM	PFOA +PFOS	10 μM (100 mM K2HPO4)	S = 0.5 cm2, anode potential = 3.6 V/SHE, pH = 7	>99.9%	morphology control	[148]
	Ce-doped Ti4O7: (Ti1–xCex)4O7	PFOS	20 nM (10 mM Na2SO4)	S = 9 cm2, L = 5 mm, J = 20 mA·cm−2, pH = 7	>83.3% (2 h)	metal-doping	[121]
	Pd/Ti4O7	PFOA	0.12 mM (50 mM Na2SO4)	S = 25 cm2, V = 30 mL, J = 10 mA·cm−2, T = 25 ± 1 °C, pH = 7.2	>90% (1 h)	metal-doping	[142]
	Mixed metal oxide (MMO)	PFOA +PFOS	5 mg L−1 (500 mg/L Na2SO4)	S = 100 cm2, L = 16 mm, J = 10 mA·cm−2, pH = 7.4	>90% (8 h)	morphology control	[149]
	LaNixY1−xO3 (Y = Fe/Cu/Co/Sr)	PFOA	0.25 mM (0.05 M Na2SO4)	L = 3 cm, J = 20 mA·cm−2, pH = 5, 1.0 mM FeSO4	90% (Y = Sr, 2.5 h)	crystal facet control	[150]
Other composites	Mxene-based membrane	PFBA +PFOA	1 μg/L (0.1 M Na2SO4)	S = 12.56 cm2, L = 10 mm, J = 10 mA·cm−2, pH = 7.00 ± 0.10	>99% (3 h)	morphology control	[19]
	Co−CN2−Fe2O3	PFOA	1.9 mg L−1 (0.05 M Na2SO4)	S = 0.2475 cm2, J = 2.2 mA·cm−2, T = 25 °C, pH = 2	70% (30 min)	metal-doping +morphology control	[20]
	Fe−Mn/CA	PFOA	50 mg mL−1 (50 mM Na2SO4)	S = 7.0 cm2, L = 2 cm, J = 2.85 mA·cm−2, T = 25 °C, pH = 3	97% (4 h)	co-doping +morphology control	[143]

5. Conclusions and Perspectives

This review presents the latest research on photo/electro-catalytic advanced oxidation processes for the degradation of PFASs. It extensively analyzes the mechanisms and pathways of photocatalytic and electrocatalytic degradation of PFASs. Based on this, the current research status of photocatalysts and electrocatalysts was systematically reviewed. Material modification methods for enhancing PFAS degradation in photocatalytic and electrocatalytic AOPs are summarized as follows:

(1)

For catalysts in the photocatalytic oxidation of PFAS systems, current research primarily focuses on improving catalyst activity by addressing the rapid recombination of photogenerated e⁻−h⁺. Methods include introducing surface defects or oxygen vacancies, metal-doping, heterojunction construction, and crystal facet regulation [21,26,28,39,71,74,96,101]. Furthermore, material composites and morphology regulation have been utilized to enhance reaction probabilities between PFASs and active groups, effectively enhancing the efficiency of PFAS degradation [25,28,40,104].

(2)

For catalysts in the electrocatalytic oxidation of PFAS systems, current research mainly focuses on increasing the yield of active groups through various methods [21,85,94]. Additionally, enhancing the efficiency of electron transfer and mass transfer processes is achieved through metal loading and constructing nano 3D structures [43,85,124,140]. The stability of electrodes is also improved through the construction of intermediate layers [63,145].

The challenges and prospects regarding the current state of development for these methods can be summarized as follows:

(1)

Regarding catalysts: Research on photocatalytic materials aims to develop novel catalysts through the combination of various modification strategies. The goal is to enhance catalyst adsorption efficiency, increase the concentration of active components on the material surface, and improve light utilization, thereby promoting the efficient transfer and separation of photogenerated charges on the photocatalytic material surface [28,35,71,90]. Research on electrocatalytic materials mainly focuses on developing efficient, inexpensive, and highly stable anode materials to enhance the surface activity sites of electrodes, thereby effectively enhancing PFAS degradation.

(2)

Beyond catalysts: In-depth study of other factors in advanced oxidation processes aims to enhance degradation efficiency of PFAS through system optimization. This includes adjusting the solution chemistry [75], adopting various strategies for synergistic catalytic oxidation [83], and conducting research on photo/electrocatalytic treatment technology to improve the utilization of catalyst materials and reduce treatment costs [30].

(3)

Current research on photo/electrocatalytic oxidation for PFAS degradation is predominantly based on ideal reaction systems. However, it is necessary to consider the actual wastewater environment, including pH, natural organic matters, coexisting ions, and the competition adsorption of degradation intermediates on the catalyst surface. Research on the competitive reactions of PFASs and coexisting substances with effective ROS should be conducted to promote the practical application of photo/electrocatalytic degradation of PFASs [59,151,152].