Recent advances in electrochemical decontamination of perfluorinated compounds from water: a review

Highlights

● Recent advances in the electrochemical decontamination of PFAS are reviewed.

● Underlying mechanisms and impacting factors of these processes are discussed.

● Several novel couped systems and electrode materials are emphasized.

● Major knowledge gaps and research prospects on PFAS removal are identified.

Abstract

Per- and polyfluoroalkyl substances (PFAS) pose serious human health and environmental risks due to their persistence and toxicity. Among the available PFAS remediation options, the electrochemical approach is promising with better control. In this review, recent advances in the decontamination of PFAS from water using several state-of-the-art electrochemical strategies, including electro-oxidation, electro-adsorption, and electro-coagulation, were systematically reviewed. We aimed to elucidate their design principles, underlying working mechanisms, and the effects of operation factors (e.g., solution pH, applied voltage, and reactor configuration). The recent developments of innovative electrochemical systems and novel electrode materials were highlighted. In addition, the development of coupled processes that could overcome the shortcomings of low efficiency and high energy consumption of conventional electrochemical systems was also emphasized. This review identified several major knowledge gaps and challenges in the scalability and adaptability of efficient electrochemical systems for PFAS remediation. Materials science and system design developments are forging a path toward sustainable treatment of PFAS-contaminated water through electrochemical technologies.

Key words： Perfluorinated compounds; Electrochemical approach; Working mechanisms; Impacting factor; Coupled process

1 Introduction

The per- and polyfluoroalkyl substances (PFAS) are a class of artificial chemicals composed of fully or partially fluorinated alkyl chains (at least one perfluoroalkyl moiety, C_nF_2n+1^–) with the terminal functional groups (e.g., carboxylate, sulfonate, and alcohol). They can have different chain lengths in the form of linear or branched isomers (Giesy and Kannan, 2002; Lindstrom et al., 2011). The terminal functional groups of PFAS are normally hydrophilic, while the fluorinated region is both hydrophobic and lipophobic. The C-F bonds have the exceptional chemical properties of strong bonding energy (~485 kJ/mol), high redox potential (F to F^–, E⁰ = 3.6 V), and perfect orbitals overlap (2s and 2p) (Lu et al., 2020). These features offered PFAS unique physicochemical characteristics, including oil and water repellency, good thermal stability, and excellent resistance to most chemicals (acids, bases, oxidants, and reductants) (Rahman et al., 2014; Shen et al., 2022). Such property has endowed their applications in diverse industrial and consumer products since the 1940s (Key et al., 1997). For example, it has been widely used as a surfactant, dispersant, and leveling agent for paints, lubricants, and mist suppression (Trojanowicz et al., 2018). As a result, a large amount of PFAS is released into the natural environment through direct release from special industrial activities (e.g., fire-fighting foams) and indirect sources from environmental degradation of chemical precursors, typically through atmospheric and biological degradation (Grandjean, 2018).

The presence of PFAS in human serum was first discovered in the late 1960s (Taves, 1968). Until the end of the 1990s, it was found that PFAS were ubiquitous in the human body and even measurable in the blood of polar bears, birds, and other species (Houde et al., 2006; Kelly et al., 2009). The accumulation of PFAS in human blood is primarily derived from their extensive distribution in water and atmospheric environment. It was recently reported that the mean PFAS concentration in drinking water ranged from 0.1 to 502.9 ng/L in China, and the PFAS was dominated by perfluorooctanoic acid (PFOA), perfluorobutanoic acid (PFBA), and perfluorooctanesulfonic acid (PFOS). For example, an extremely high concentration of PFOA (3165 ng/L) in drinking water was found in Zigong, China, due to a fluorochemical plant and multiple industries such as leather, textile, and paper manufacturing (Liu et al., 2021b). In addition, because the precursors of PFAS are usually volatile and easy to be adsorbed on particulate matters, high PFAS concentrations are often present in the atmosphere, although most of them can be removed by precipitation (Wang et al., 2021). Fig.1 illustrates an overview of transportation and distribution of PFAS in the natural environment.

Fig.1 Overview of the transportation and distribution of the PFAS in the natural environment.

Full size|PPT slide

However, regulators only started calling for an assessment of the toxicity of PFAS in the early 2000s (Giesy and Kannan, 2001). The tracking studies confirmed a connection between PFAS exposure and health problems, including increased cholesterol and liver enzymes, high probability of testicular and kidney cancer, decreased fertility and fecundity, immune suppression, and thyroid disorders (Steenland et al., 2010). Besides the toxicity evaluation, implementing relevant countermeasures and introducing standards have also been carried out. Between 2000 and 2002, the production of PFOS and related compounds were voluntarily phased out in the 3M Company, which was once the major PFOS producer. In 2016, the Environmental Protection Agency (EPA) of the United States established a drinking water health advisory level of 70 ng/L for a combined concentration of PFOA and PFOS (Radjenovic et al., 2020). Moreover, in the recent Basel, Rotterdam, and Stockholm Conventions, PFOS and PFOA have been revised and added to the list of Convention on POPs in Annex B (Restriction of PFOS and its salts) and Annex A (Elimination of PFOA and its salts) (Lu et al., 2020). Although many actions so far have exerted a positive effect on reducing the exposure to PFAS, some processes (e.g., fire-fighting foams, semiconductors, and metal plating) still have specific exemptions (Hogue, 2019). It has been estimated that the accumulated emission of C4–C14 PFAS will be about 20–6420 t between 2016 and 2030 (Wang et al., 2014b), suggesting the protection of drinking water sources from PFAS contamination is urgently needed.

The remediation of PFAS from water has become one of the most extensively investigated subjects. The rapid development in this field is reflected by the increasing number of research papers and review articles published in the last decade. Current PFAS treatment strategies can be divided into several categories: adsorption, membrane filtration, photochemical, electrochemical, and other advanced oxidation processes. Among them, the electrochemical strategy is promising for PFAS decontamination because of its desirable attributes like free of chemical addition, less secondary pollution, mild reaction conditions, and easy automation (Martinez-Huitle et al., 2015; Nzeribe et al., 2019; Kim et al., 2020b; Pierpaoli et al., 2021; Saeidi et al., 2021; Hwang et al., 2022; Veciana et al., 2022). Two recent reviews on the evaluation of PFAS destruction in electro-oxidation processes were reported. Radjenovic et al. (2020) provided important insights into the mechanisms of PFAS degradation and the effects of water chemistry on PFAS degradation during electro-oxidation. Meanwhile, Sharma et al. (2022) reviewed the recent development of anode materials for the electro-oxidation of PFAS. On the other hand, progress on the electro-oxidative degradation of PFAS is rapidly evolving. In particular, several coupled systems have been recently developed to boost the PFAS removal efficacy with reduced energy consumption. Except for electro-oxidation, electro-adsorption, and electro-coagulation, two other major electrochemical technologies, have also been widely reported for PFAS removal (Wang et al., 2014a; Mu et al., 2021; Shrestha et al., 2021; Li et al., 2022b). However, there is no scientific review of these important advances.

Here, we aim to provide a comprehensive update focusing on the underlying working mechanisms of electrochemical strategies for electro-oxidation, electro-adsorption, and electro-coagulation in PFAS decontamination. Meanwhile, we attempted to provide key know-how on the impact of water chemistry, operational parameters, and reactor configurations on the PFAS removal performance. In addition, we highlighted the recent advancements in several coupled systems to enhance the PFAS electro-oxidative degradation. We also identified the main research limitations of electrochemical technologies and put forward some possible important research needs in the future.

2 Electro-oxidation technology

2.1 Electro-oxidation mechanisms of PFAS

The electro-oxidation of PFAS involves the electron transfer from PFAS to the anode or the hydroxide radical (HO^•)-mediated oxidation, namely, direct or indirect electrochemical oxidation (Fig.2(a)) (Niu et al., 2016). Among them, it is generally accepted that the direct electron transfer process is the rate-limiting and primarily responsible for initiating the electro-oxidation of PFAS because the HO^• is ineffective for the PFAS decomposition, which was evidenced by both experimental study and theoretical calculations (Carter and Farrell, 2008; Trojanowicz et al., 2018). In recent years, many efforts have been devoted to exploring the degradation mechanisms of PFAS during electro-oxidation. DFT calculations and reaction intermediates identification have established three degradation pathways (Niu et al., 2013; Radjenovic et al., 2020; Duinslaeger and Radjenovic, 2022), which are illustrated in Fig.3.

Fig.2 Schematic illustrations of (a) electro-oxidation, (b) electro-adsorption and (c) electro-coagulation for PFAS removal.

Full size|PPT slide

Fig.3 Proposed main reaction pathways for electro-oxidation of PFAS.

Full size|PPT slide

For PFOA (C_nF_2n+1COO^–) and PFOS ( {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}SO₃^–), the –COO^– or –SO₃^– functional group initially transfers an electron to the anode to form the unstable C_nF_2n+1COO^• and C_nF_2n+1SO₃^•, respectively (Eq. (1)), which only occurs at high anodic potentials (i.e., > 3.0 V vs. SHE) (Niu et al., 2013). They undergo the Kolbe decarboxylation or desulfonation reactions to generate C_nF_2n+1^• (Eq. (2)) (Lin et al., 2012). Subsequently, the C_nF_2n+1^• can react with HO^•, O_2, or H₂O to produce fragmented PFAS, CO_2, and HF via three parallel reaction pathways (Three Cycles) (Niu et al., 2013).

(1)\begin{array}{rl}& {\mathrm{C}}_n{\mathrm{F}}_{2n+1}\mathrm{C}\mathrm{O}{\mathrm{O}}^{-}/{\mathrm{C}}_n{\mathrm{F}}_{2n+1}{\mathrm{S}\mathrm{O}}_3^{-}\to \\ & {\mathrm{C}}_n{\mathrm{F}}_{2n+1}\mathrm{C}\mathrm{O}{\mathrm{O}}^{\bullet }/{\mathrm{C}}_n{\mathrm{F}}_{2n+1}{\mathrm{S}\mathrm{O}}_3^{\bullet }+{\mathrm{e}}^{-}\end{array}

(2){\mathrm{C}}_n{\mathrm{F}}_{2n+1}\mathrm{C}\mathrm{O}{\mathrm{O}}^{\bullet }/{\mathrm{C}}_n{\mathrm{F}}_{2n+1}{\mathrm{S}\mathrm{O}}_3^{\bullet }\to {\mathrm{C}}_n{\mathrm{F}}_{2n+1}^{\bullet }+\mathrm{C}{\mathrm{O}}_2/\mathrm{S}{\mathrm{O}}_3

For the first path (Cycle I), the {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}^• can react with the HO^• generated from the anodic oxidation of water to form C_nF_2n+1OH (Eq. (3)), which can further react with HO^• to generate C_nF_2n+1O^• (Eq. (4)). Then, the C_nF_2n+1O^• would spontaneously release the CF₂O and thus produce a shorter C_n–1F_2n+1^• (Eq. (5)) by eliminating a CF₂ unit. The generated C_nF_2n+1^• can be further destructed repeatedly via similar steps until the multiple CF₂ units are gradually lost. This degradation pathway was supported by experimental results, such as the electrolysis of PFOS at a boron-doped diamond (BDD) anode. Only SO₄^2– and F^– were produced without other intermediate products (Carter and Farrell, 2008; Trautmann et al., 2015).

(3){\text{C}}_n{\text{F}}_{2n+1}^{\bullet }+{\text{HO}}^{\bullet }\to {\text{C}}_n{\text{F}}_{2n+1}\text{OH}

(4){\text{C}}_n{\text{F}}_{2n+1}\mathrm{O}\mathrm{H}+{\text{HO}}^{\bullet }\to {\text{C}}_n{\text{F}}_{2n+1}{\text{O}}^{\bullet }+{\text{H}}_2\text{O}

(5){\text{C}}_n{\text{F}}_{2n+1}{\text{O}}^{\bullet }\to {\text{C}}_{n-1}{\text{F}}_{2n-1}^{\bullet }+{\text{CF}}_{\text{2}}\text{O}

The second route (Cycle II) also generates {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}OH first through the reaction between {\text{C}}_{\text{n}}{\text{F}}_{\text{2}\text{n}\text{+1}}^• and HO^• (Eq. (3)). Then, the thermally unstable {\text{C}}_{\text{n}}{\text{F}}_{\text{2}\text{n}\text{+1}}OH undergoes intramolecular rearrangement to eliminate HF and to form the more stable C_n–1F_2n–1COF (Eq. (6)) (Bentel et al., 2019), which would further react with H₂O to form C_n–1F_2n–1COO^– (Eq. (7)). So far, one CF₂ unit has been removed. Repeating the cycle steps allows PFAS to be completely mineralized to CO₂ and HF (Zhuo et al., 2012; Liu et al., 2019b; Duinslaeger and Radjenovic, 2022). However, DFT calculations suggested that the release of HF had a high activation barrier (i.e., 223 kJ/mol), implying that this route played a minor role in the electro-oxidation system (Niu et al., 2013).

(6){\text{C}}_n{\text{F}}_{2n+1}\mathrm{O}\mathrm{H}\to {\text{C}}_{n-1}{\text{F}}_{2n-1}\mathrm{C}\mathrm{O}\mathrm{F}+\mathrm{H}\mathrm{F}

(7){\text{C}}_{n-1}{\text{F}}_{2n-1}\mathrm{C}\mathrm{O}\mathrm{F}+{\text{H}}_{\text{2}}\mathrm{O}\to {\text{C}}_{n-1}{\text{F}}_{2n-1}\text{CO}{\text{O}}^{-}+\mathrm{H}\mathrm{F}+{\text{H}}^{+}

The third degradation pathway (Cycle III) involves O₂ that reacts with {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}^• to form {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}OO^• (Eq. (8)). The {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}OO^• can react with another R_FCOO^• (e.g., {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}OO^•) to produce {\text{C}}_n{\text{F}}_{\text{2}n\text{+1}}O^• (Eq. (9)), which decays to form C_n–1F_2n–1^• with one CF₂O removed (Eq. (10)) (Pierpaoli et al., 2021; Hwang et al., 2022). By repeating the process, the C_n–1F_2n–1^• can continue the chain shortening through Cycles I–III. Of note, this pathway describes an HO^•-free reaction mechanism, which may be the main electro-oxidation pathway for PFAS with low HO^• concentration or in the presence of HO^• scavengers.

(8){\text{C}}_n{\text{F}}_{2n+1}^{\bullet }+{\text{O}}_2\to {\text{C}}_n{\text{F}}_{2n+1}{\text{OO}}^{\bullet }

(9){\text{C}}_n{\text{F}}_{2n+1}{\text{OO}}^{\bullet }+{\text{R}}_{\text{F}}\text{C}{\text{OO}}^{\bullet }\to {\text{C}}_n{\text{F}}_{2n+1}{\text{O}}^{\bullet }+{\text{R}}_{\text{F}}\text{C}{\text{O}}^{\bullet }+{\text{O}}_{\text{2}}

(10){\text{C}}_n{\text{F}}_{2n+1}{\text{O}}^{\bullet }\to {\text{C}}_{n-1}{\text{F}}_{2n-1}^{\bullet }\text{+}{\text{CF}}_{\text{2}}\text{O}

Besides the three reaction pathways mentioned above, the sulfate radical (SO₄^•–) may also be formed by the oxidation of SO₄^2– at a high anodic potential, which may also contribute to the degradation of PFAS (Liu et al., 2019a; Liu et al., 2019b). For example, when using a B/N co-doped diamond as the anode with an anodic current density of 4.0 mA/cm², the degradation rate of PFOA (50 mg/L) was 2.4–3.9 folds higher in the Na₂SO₄ electrolyte than those in the NaNO₃ or NaClO₄ electrolytes (Liu et al., 2019b). For the SO₄^•– mediated oxidation system, the degradation mechanisms of PFAS and the influence factors (e.g., solution pH, S₂O₈^2– dose, and water matrix constituents) have been summarized and discussed in detail in a published review paper (Yang et al., 2020).

To obtain a larger amount of HO^• with improved the degradation of PFAS, the electrochemical process with the synergistic cathodic electro-Fenton and anodic oxidation has been proposed (Liu et al., 2015; Wang et al., 2019a). The PFAS first transferred one electron to the anode to form the PFAS radical, followed by the decarboxylation or desulfonation to produce C_nF_2n+1^•. The highly active C_nF_2n+1^• can react not only with the HO^• formed at the anode through water oxidation but also with the additional HO^• generated at the cathode via an electro-Fenton process (Fig.4), which significantly increases the PFAS degradation efficiency. Liu et al. (2015) reported a synergic system for efficient mineralization of PFOA (50 mg/L) when a BDD (anode) and hierarchical porous carbon (cathode) were used as the active electrodes. A total organic carbon removal efficiency of 70.4 %–90.7 % could be achieved within 4 h at a low potential of 0.4 V in a wide pH range (2–6). Meanwhile, the current efficiency for PFOA degradation was one order of magnitude higher than that of pure anodic electrochemical oxidation. Coincidentally, Wang et al. (2019a) also showed that the destruction of PFOA could reach 97 % within 4 h using BDD as the anode and FeMnC as the cathode, respectively. These reports suggest that the cathode material and the associated reaction at the cathode are also important for PFOA degradation.

Fig.4 Schematic illustration of degradation mechanism of PFAS by anodic electro-oxidation and cathodic electro-Fenton.

Full size|PPT slide

2.2 Factors influencing the electron-oxidation performance of PFAS

2.2.1 Operating conditions

1) Current density

Current density is the most important parameter determining the yield of HO^• as well as the electron transfer rate from the PFAS on the electrode surface (Yang et al., 2022). In general, the PFAS degradation is proportional to the applied voltage. Shi et al. (2019) employed the porous Magnéli phase titanium suboxide (Ti₄O₇) ceramic membrane as the anode and found that the degradation of PFOS increased from ~20 % to ~98 % with the increase of the current density from 1.0 to 4.0 mA/cm². However, the current density usually has an optimal value, beyond which the degradation of PFAS will stagnate or even decline. This is primarily attributed to the occurrence of an oxygen evolution side reaction that inhibits the approaching of PFOA molecules toward the anode surface. For example, a PFOA degradation efficiency of 96.1 % could be obtained with a current density of 20 mA/cm² on a Zr-PbO₂ anode, while the efficiency was only increased by 0.9 % when further increasing the current density to 30 mA/cm² (Xu et al., 2017). High current density also increases energy consumption and shortens the service life of the electrode.

2) Solution pH

The solution pH affects the yield of HO^•, oxygen evolution potential, and the service life of the electrode. Thus, it is necessary to evaluate its effects on the degradation of PFAS (Lin et al., 2012; Zhuo et al., 2012). There is no reported consensus on the effects of pH to date. Some studies claimed that acidic conditions could facilitate the degradation of PFAS in the electro-oxidation system, while others showed that PFAS degradation was not affected by the solution pH. For instance, Lin et al. (2012) reported that the degradation rate of PFOA at pH 5.0 was approximately twice as high as that at pH 11 by a Ti/SnO₂-Sb anode. This could be attributed to the inhibited oxygen evolution under the acidic condition and the reduced competitive adsorption of OH^– on the anodic surface (Lin et al., 2012). Nevertheless, another study exhibited that the oxidation rate of PFOA on a Ti₄O₇/Pd anode only changed marginally across a broad pH range from 5.2 to 9.1 (Huang et al., 2020). The local environment near the anode surface may become more acidic than the bulk pH when the high potential is applied, which will significantly affect the degradation of PFAS (Lin et al., 2020). If only relying on the initial pH for its effects on the degradation of PFAS, it is likely to lead to a false interpretation. Thus, real-time monitoring of the pH near the anode surface is necessary. Furthermore, it is worth noting that the pH may also affect the wettability and electrical conductivity of the electrode surface, which can also affect the degradation of PFAS (Radjenovic et al., 2020). Overall, it is difficult to extrapolate the reported information on the effects of pH on PFAS electro-oxidation to a single mechanism.

3) Electrolyte type

The electrolyte offers a conductive medium for ions with minimized voltage drop and the resistance of an electrochemical system (Song et al., 2010). Different electrolyte types have different conductivities, which can influence the thickness of the electric double layer at the electrode surfaces. The increase in the double layer thickness reduces the ion transport toward the electrode surface and the electro-oxidative activity, thereby deteriorating the degradation performance of PFAS (Radjenovic et al., 2020). In addition, increased electrolyte viscosity and surface tension may decrease the electrode wettability and thus also may inhibit the electro-oxidative degradation of PFAS (Wu et al., 2004). Zhuo et al. (2014) compared the effects of three electrolytes on the degradation of 6:2 fluorotelomer sulfonic acid (20 mg/L) on a Ti/SnO₂-Sb₂O₅-Bi₂O₃ anode. They found that its degradation declined along with the trend of the electrolytes NaClO₄ > NaCl > Na ₂SO₄. Barisci et al. (2021) also showed a higher removal of PFOA and PFOS (100 μg/L) in NaCl than in a Na₂SO₄ electrolyte. Linear sweep voltammetry curves showed that the oxygen overpotential on the Ti/RuO₂ anode was higher in the NaCl electrolyte, suggesting its higher oxidation potential resulted in the increased degradation efficiency, although the presence of halide ions may lead to the formation of toxic byproducts (e.g., Cl₂, HClO, and ClO_x^–) during the electrochemical oxidation (Azizi et al., 2011; Yang et al., 2019). This risk needs to be mitigated, especially when high electrolyte concentration and high potential are employed in the electro-oxidation systems. Due to the limited publications, it is difficult to identify the general effect of electrolyte type on PFAS degradation based on the currently limited data.

2.2.2 Anode materials

1) Boron doped diamond (BDD) electrode

The biggest obstacle to the high-efficiency electro-oxidation of PFAS is the occurrence of oxygen evolution reaction (OER). An anode material with a high oxygen evolution potential is required to inhibit the undesirable OER. BDD electrode has been widely used as the anode for the PFAS treatment due to its high conductivity and wide working window for the OER (Pierpaoli et al., 2021; Uwayezu et al., 2021; Liang et al., 2022a). Meanwhile, the BDD electrode presents excellent electron transfer ability, superior chemical stability, longer service life, and mechanical strength, which are essential for the industrial application of electrochemical treatment to prevent scaling and fouling (Radjenovic et al., 2020). The weak physical adsorption of HO^• and other oxygen species on the BDD surface also promotes the interaction between PFAS and the active sites on the anode surface (Carter and Farrell, 2008). Uwayezu et al. (2021) showed that the BDD anode could achieve a high PFOA degradation (99.5 %) at a high current density (21.4 mA/cm²) and long oxidative time (4 h). Some strategies have been proposed to improve the electrocatalytic performance, including N doping in BDD or introducing the polarized carbon atoms sites acting as the reactive sites (Liu et al., 2019b; Barisci and Suri, 2020; Pierpaoli et al., 2021; Nienhauser et al., 2022). For instance, compared to BDD and N-doped diamond, the oxidation and mineralization kinetics of PFOA were increased by 2.0–2.4 times using the N-doped BDD electrode (Liu et al., 2019b).

Nevertheless, the high cost of chemical deposition (~7000 $/m²) (Chaplin, 2019) and the limitations of low surface area (< 10 cm²) have greatly hindered the use of BDD electrodes (Radjenovic et al., 2020). In addition, the lack of substate for the diamond coating further limits the large-scale production of BDD electrodes. Although previous studies demonstrated that Ta, Nb, and W substrates could offer brilliant chemical and mechanical stability with high electrical conductivity, they are limited by excessive material cost (Radjenovic et al., 2020). Recently, Si or Ti substrates have also been used, showing excellent PFAS degradation (Lin et al., 2018; Barisci and Suri, 2020). Unfortunately, Si is very brittle with poor electrical conductivity, while Ti is also limited by its short service life. Therefore, it remains a major challenge to synthesize an inexpensive, high specific area, excellent conductivity, and long service life electrode for highly efficient PFAS degradation.

2) Mixed metal oxide electrode

Due to the high cost of BDD electrodes, mixed metal oxides (MMOs), primarily based on SnO₂ and PbO₂ coatings, have also been produced as alternative anode materials (Zhuo et al., 2011; Niu et al., 2012; Zhao et al., 2013). Pristine SnO₂ cannot be used directly as an anode due to its high resistance and poor stability (Wang et al., 2020a). Hence, dopants, such as Sb, F, and IrO₂, are usually introduced to improve electrical conductivity and corrosion resistance (Lin et al., 2013; Yang et al., 2015). Lin et al. (2012) demonstrated the degradation and defluorination of PFOA (100 mg/L) over the Ti/SnO₂-Sb anode at10 mA/cm² with the efficiencies of 98.8 % and 73.9 %, respectively, at pH 5.0 in 90 min. However, the passivation of the Ti substrate and the formation of a nonconductive Sn hydroxide layer resulted in the short service life of the Ti/SnO₂-Sb anode (Santos et al., 2014). Introducing corrosion-resistant additive metals, such as Bi and Ce, can effectively suppress surface passivation (Zhuo et al., 2011). For instance, the doping of Bi made the electrode surface denser, which prevented the diffusion of O₂ into the Ti substrate with a reduced rate of TiO₂ passivation (Zhuo et al., 2011). The accelerated service life test indicated the doubling of the service life of the Ti/SnO₂-Sb anode (i.e., 0.8 h vs. 0.4 h) with the Bi doping. Some studies used carbon materials, including carbon aerogel (CA) and CNT), to replace the Ti substrates for the SnO₂-Sb electrode, which allows the uniform distribution of SnO₂-Sb on the substrates (Zhao et al., 2013). Zhao et al. (2013) compared the CA/SnO₂-Sb electrode and Ti/SnO₂-Sb electrode and found that the defluorination of PFOA on the CA/SnO₂-Sb electrode was 6 times higher than that on the Ti/SnO₂-Sb electrode.

PbO₂ is another promising non-active anode material for PFAS treatment because of its low cost and high electrical conductivity (Niu et al., 2012; Duan et al., 2020). PbO₂ usually presents needle-like, and bulk crystal structures named α-PbO₂ and β-PbO₂, respectively. Similar to the SnO₂ anode, Ti still is the common substrate material for the PbO₂ electrode, but the interface resistance between PbO₂ and Ti substrate is large, which may lead to the shedding of the PbO₂ layer. The insertion of an intermediate layer (e.g., SnO₂-Sb and TiO₂ nanotube array) between the Ti substrate and the PbO₂ layer can effectively alleviate this problem (Zhong et al., 2013; Zhuo et al., 2016). It was shown that an excellent performance with 91.1 % degradation and 77.4 % defluorination of PFOA was achieved using a Ti/SnO₂-Sb/PbO₂ anode (Lin et al., 2012). In addition, incorporating dopants (e.g., Ce, Co, Bi, and F) into the Ti/SnO₂-Sb/PbO₂ anode could further increase their physical stability and electrochemical reactivity (Zhuo et al., 2011). For instance, the unique outer electronic structure of the Ce element could increase the electron transfer efficiency of the electrode and offer new nucleation sites for the PbO₂, which can influence the average grain size of the PbO₂ (Lin et al., 2013).

3) Titanium suboxides (Ti₄O₇) electrode

Recently, the Magnéli phase Ti₄O_7, with high oxygen evolution potential (> 2.5 Vvs. SHE), good chemical inertness, and large specific surface area, has attracted some attention as a favorable anode material for the oxidation of PFAS (Geng et al., 2015; Lin et al., 2018; Wang et al., 2020b). A large number of oxygen vacancies in the Ti₄O₇ also endow its excellent electrochemical activity with good electrical conductivity (1000 S/cm) (You et al., 2016). For example, nearly 100 % of PFOA was degraded on Ti₄O₇ ceramic anode within 2 h in a batch mode (Lin et al., 2018). Recently, Wang et al. (2022) also demonstrated that the nano-Ti₄O₇ anode outperformed the micro-Ti₄O₇ anode in terms of the reaction rates and energy efficiency for the PFOS degradation, suggesting that the PFAS degradation was strongly dependent on the pore size of Ti₄O₇. However, compared to the traditional inactive anode materials such as BDD and MMO, the pure Ti₄O₇ still exhibits a lower electron transfer rate, resulting in a low yield of HO^• (Chaplin, 2014; Lin et al., 2021). For example, a study showed that the chemical yield of 7-hydroxycoumarin (a probe for HO^•) generated by the oxidation of coumarin at Ti₄O₇ anode (~40 %) was higher than that at BDD anode (< 5 %) within 150 min, indicating that BDD anode produced more HO^• than Ti₄O₇ anode (Bejan et al., 2012). To date, few strategies have been proposed to improve the oxidation of PFAS at the Ti₄O₇ electrode. Doping metal or carbon can regulate the surface properties and electronic structures of Ti₄O₇ and thus promoting its electrochemical reactivity (Le et al., 2019; Huang et al., 2020; Lin et al., 2021). For instance, the introduction of amorphous Pd clusters can enhance the electron transfer via Pd-O bonds, which resulted in a 5-fold increase in the oxidation kinetics of PFOA (~2.02 h^–1) compared to the pristine Ti₄O₇ electrode (~0.41 h^–1) (Huang et al., 2020). In another study, a Ce³⁺-doped Ti₄O₇ electrode exhibited an increased yield of HO^• (37 %–129 %) since the variable electronic states of Ce were beneficial to the formation of oxygen vacancy on the Ti₄O₇ surface (Lin et al., 2021). A recent first principles study of the PFAS degradation on a Ti₄O₇ electrode confirmed the effectiveness of the doping strategies (Li et al., 2021a). On the other hand, the yield of HO^• can be improved by adding oxidants. The use of peroxymonosulfate (PMS) in the electro-oxidative process with Ti₄O₇ electrode could achieve 100 % degradation of PFOS within 40 min because the PMS activation could boost HO^• generation during electrochemical process (Li et al., 2022a). Notably, there is no systematic study to evaluate the differences between these popular electrode materials (e.g., BDD, PbO₂, and Ti₄O₇) in degrading PFAS under identical conditions, although such knowledge is beneficial to the future development of anode materials.

2.2.3 Reactor configuration

The design and configuration of a reactor directly determine the efficiency of mass transferring in an electrochemical system, which may significantly affect the PFAS degradation kinetics and efficiency. Increasing the mass transfer to the electrode surface will improve the electrochemical reaction rate and increase the current efficiency with reduced energy consumption (Ren et al., 2022). Conventional rod-like or block electrodes in the batch system mainly rely on mass diffusion, which greatly reduces the reaction kinetics (Fig.5(a)). In contrast, a reactive electrochemical membrane (REM) design may provide better electro-oxidation activity due to the enhancement of mass transport by convection (Liu et al., 2020). Shi et al. (2019) systematically compared the degradation of PFOS on an Ebonex anode by batch and REM systems and found that the mass-transfer rate constant was increased from 2.56 × 10^–6 to 6.21 × 10^–6 m/s.

Fig.5 Schematic illustrations of electrochemical reactor operated in (a) batch, (b) flow-by, and (c) flow-through modes.

Full size|PPT slide

In addition, employing the REM in the flow-by or flow-through mode also significantly affects the PFAS degradation rate. The flow-by mode usually results in the formation of a thick diffusional boundary layer on the REM surface (~100 μm) (Fig.5(b)). Thus, the PFAS reaction rate is still diffusion-limited at low current densities (Chaplin, 2019). By contrast, the PFAS degradation rate will be significantly increased using a flow-through system due to the convective transportation of the solution through the small pores in the electrode, which results in effective exposure of the electrode surface (Fig.5(c)) (Chaplin, 2019). It has been reported that the electro-oxidation mass transfer rate constants of PFOA and PFOS in a flow-through Ti₄O₇ REM system were 1–2 orders of magnitude higher than those of the flow-by mode (i.e., 4.4 × 10^–5 vs. 5.74 × 10^–6 m/s for PFOA, 1.3 × 10^–4 vs. 8.81 × 10^–6 m/s for PFOS) (Le et al., 2019). These promising results were obtained under favorable electrochemical conditions with high solution conductivity at a high initial PFAS concentration. Therefore, it is necessary to study the performance in real waste streams to evaluate the feasibility of REM treatment strategy on an industrial scale.

3 Electro-adsorption technology

3.1 Electro-adsorption mechanisms of PFAS

Electro-adsorption technology, also known as capacitive deionization technology, is a common water treatment process that combines the adsorption separation process with electrochemistry (Huang et al., 2014). Electro-adsorption technology is a non-Faraday process that does not involve electron gain or loss. The current needed in the electro-adsorption process is only for maintaining a double electric layer. Therefore, the energy consumption and the operation cost are low (Ma et al., 2016). Compared to the traditional adsorption method, electro-adsorption can improve the adsorption capacity and rate due to electrostatic attraction, and achieve the regeneration of electrode material by applying a reverse potential (Kim et al., 2020a; Shrestha et al., 2021).

Electro-adsorption technology has been widely used to remove PFAS (Saeidi et al., 2021; Tian et al., 2021). Carbon materials, such as activated carbon, carbon aerogels, graphene, and carbon nanotubes, are the preferred electrode materials for electro-adsorption (Bayram and Ayranci, 2010; Luo et al., 2018). Nevertheless, neither the high electron transfer resistance of activated carbon nor the microporous structure of carbon aerogel is conducive to good electro-adsorption performance (Li et al., 2011). On the other hand, graphene and multi-walled carbon nanotubes (MWCNTs) present excellent electrochemical stability with conductivity. They are expected to be excellent electrode materials for the electro-adsorption of PFAS in water (Wang et al., 2014a; Niu et al., 2017).

The underlying working mechanism of an electro-adsorption process is that, under the action of the applied electric field, the negatively charged PFAS in the solution migrate directionally toward the anode. Hence, PFAS are enriched on the anode surface due to the electrostatic attraction and complexation, which accelerates the removal of PFAS (Fig.2(b)). For example, compared to the suspended MWCNTs powder (0.9 × 10^–3 mmol/(h·g) and 0.0065 mmol/g), the initial adsorption rate (54.5 × 10^–3 mmol/(h·g)) and the adsorption capacity (0.98 mmol/g) of MWCNT electrode for PFOA (100 μg/L) at 0.6 V were dramatically increased by 60-fold and 150-fold, respectively (Li et al., 2011). Notably, the electro-adsorption processes involve not only the electrostatic interaction but also other adsorption mechanisms (e.g., hydrophobic interaction, F-F interaction, and hydrogen bonding) by modifying the electrode materials (Liu et al., 2018a; Quan et al., 2020). Liu et al. (2021a) prepared a reduced graphene oxide aerogel electrode modified by Cu nanoparticles and fluorine (Cu/F-rGA). They found that the PFOA removal capacity by the Cu/F-rGA electrode was 45.9 % higher at 0.8 V than that of the unmodified rGA electrode. This could be attributed to the ligand exchange between the PFOA and the Cu nanoparticles on the surface and the F-F interaction between the F dopant and the PFOA. Kim et al. (2020a) synthesized a polymer-functionalized CNT electrode containing the groups of amines (N-H) and nitroxide radicals (N-O^•). The N-H groups have served as attractive binding sites for the PFAS, and the redox-active N-O^• can be charged to form the oxoammonium cation (+N=O). Results indicated that the electrode had the ultrahigh PFOA (1 mmol/L) adsorption capacity (> 1000 mg/g) at 1.0 V since the PFOA were bound on both the N-H and the +N=O sitesvia electrostatic attraction (Fig.6(a)) (Kim et al., 2020a). More importantly, the author found that the bound PFOA could be effectively desorbed due to the charge repulsion when a negative bias was applied (e.g., regeneration of > 80 % at –1.5 V) ( Fig.6(b)). This study provided an insight into how to enhance the PFAS removal and electrode regeneration during electro-adsorption processes.

Fig.6 A schematic illustration showing electrochemically-controlled (a) capture and (b) release of PFOA by polymer-functionated CNT electrode. Adapted from (Kim et al., 2020a).

Full size|PPT slide

3.2 Effect of operating conditions on the electro-adsorption performance of PFAS

3.2.1 Applied voltage

Generally, the PFAS adsorption performance of an electrode increases with the increase of the applied voltage (Wang et al., 2014a; Liu et al., 2021a). This is mainly due to the increase in the thickness of the electrical double layer formed at the interface between the electrode and the solution, which can effectively improve the adsorption capacity of the electrode (Haro et al., 2011). For example, the PFOA adsorption capacity on a Cu/F-rGA anode was increased from 3.46 to 6.31 mg/g with an increase of bias from 0 to 0.8 V (Liu et al., 2021a). Wang et al. (2014a) observed a similar phenomenon in a continuous flow mode with an MWCNT electrode and found that the removal of PFOS (64.9 %) was higher than that of PFOA (52.0 %) under the same conditions. This was because PFOS with a longer perfluoroalkyl chain can obtain a stronger hydrophobic interaction (Li et al., 2011). Although a better PFAS removal can be achieved at a higher applied voltage, choosing the voltage below the water hydrolysis voltage of 1.23 V is recommended to reduce energy waste (Wang et al., 2014a).

3.2.2 Electrolyte concentration

Currently, the electrolytes used in the electro-adsorption system are usually Na₂SO₄ or NaCl (Kim et al., 2020a; Liu et al., 2021a), and the removal of PFAS can be optimized at a specific electrolyte concentration. Although the increase of electrolyte concentration improves the current density and accelerates the migration of PFAS, once it exceeds the critical level, inorganic anions (e.g., SO₄^2–) rather than the PFAS target will be driven in the electrolyte and occupy the adsorption sites on the electrode, resulting in the decrease in the adsorption of PFAS (Li et al., 2011). For instance, the PFOA removal by the Cu/F-rGA anode was enhanced from 5.65 to 6.38 mg/g when the concentration of Na₂SO₄ was increased from 0 to 1.0 mmol/L. However, this was decreased with further increasing the Na₂SO₄ concentration (e.g., 20 mmol/L, 4.09 mg/g) (Liu et al., 2021a). In another study, a redox-copolymer electrode was employed for the efficient removal of hexafluoropropylene oxide dimer acid (GenX) at 0.8 V (vs. Ag/AgCl) using NaCl electrolyte (Medina et al., 2021). Results showed that the optimal adsorption capacity could be obtained when the NaCl concentration was 10 mmol/L. A higher Cl^– concentration may screen more of the positive charges on the polymer, thus weakening its attractive electrostatic interactions for the GenX adsorption. Other reports also investigated the influence of the electrolyte concentration on the removal of PFAS in the electro-adsorption system and obtained similar conclusions (Li et al., 2011; Wang et al., 2014a).

3.2.3 Solution pH

Like the electro-oxidation technology, the current reports on the effect of initial solution pH are not completely consistent. The solution pH change with time is negligible since a small voltage was applied in the electro-adsorption process (Liu et al., 2021a). Some studies indicated that the initial pH does not affect the electro-adsorption of PFAS since it is dominated by the action of the applied electric field. For example, the PFOA electro-adsorption on the Cu/F-rGA electrode yielded similar removal efficiency from pH 4.2 (95.61 %) to 10.1 (88.47 %) (Liu et al., 2021a). Conversely, several studies found that the initial pH significantly affected the electro-adsorption of PFAS. Niu et al. (2017) reported that the removal of PFOA on a CNTs-20 % graphene composite electrode was increased by 52 % as the pH was increased from 3.0 to 9.0. PFOA will be more negatively charged at pH 9.0 since its pKa is about 2.8. Thus, high pH could enhance the electrostatic attraction between PFOA and the anode. Yet, Saeidi et al. (2021) demonstrated an opposite phenomenon: the removal of PFOA (1 mg/L) at pH 3.0 on activated carbon felts was higher than at pH 9.0. This was because the carbon surface (pKa of 7.1) was more negatively charged at pH 9.0, which would increase the electrostatic repulsion to the negatively charged PFOA and thus reduce the adsorption capacity of the electrode.

These studies seem to provide some plausible explanations for the experimental results. Unfortunately, the pH dependence electro-adsorption behavior depends on the polarization of the surface charge for specific material as well as the charge on the PFAS. However, many of these reports failed to consider the coherent effects of the pH on both the electrode and the PFAS state. Thus, to ensure the accuracy of the interpretation of the experimental results, the effects of pH on the electro-adsorption of PFAS should be explored as comprehensively as possible, although it is still difficult.

3.2.4 Plate distance

Enhanced PFAS removal can be obtained by narrowing the distance between the cathode and anode. Smaller electrode spacing can produce a stronger electrical field near the electrode, forming a thicker double electric layer. This shortens the time for the negatively charged PFAS to reach the polarized anode surface, thus enhancing the adsorption performance of the electrode (Niu et al., 2017). An experimental study showed that the removal of 100 μg/L PFOA after 60 min by a CNTs-graphene composite electrode was 94.6 %, 82.7 %, and 60.6 % for the electrode separation distance of 5, 10, and 15 cm, respectively (Niu et al., 2017). Nevertheless, in the conventional parallel planar electrode system, when the plate spacing is too small, it may result in low throughput. This drawback may be significantly improved when the system is operated in a flow-through mode, which needs to be demonstrated in the follow-up studies.

4 Electro-coagulation technology

4.1 Electro-coagulation mechanisms of PFAS

Electro-coagulation technology has been widely concerned and successfully demonstrated for PFAS remediation (Bao et al., 2020; Opoku-Duah and Johnson, 2020; Mu et al., 2021; Li et al., 2022b). Compared to traditional coagulation technology, the electro-coagulation process presents several advantages in terms of lower sludge yield, no addition of chemical coagulant or flocculants, simple equipment, and ease of operation (Kim et al., 2020b; Qi et al., 2020). Electro-coagulation technology usually uses soluble metal materials such as Fe and Al as the anode. Under the external electric field, many cations (e.g., Fe²⁺, Fe³⁺, and Al³⁺) can be produced from the metal anode (Si et al., 2021). These cations can further generate various amorphous hydroxyl complexes or hydroxide flocs by a series of reactions, including hydrolysis and polymerization. The PFAS in water can be removed by coagulation and adsorption (Fig.2(c)) (Mollah et al., 2004). During this process, H₂ and O₂ bubbles may be generated at the cathode and anode. They can be adsorbed on the flocs. The flocs thus will be brought to the surface of the solution through electrical floatation to achieve solid-liquid separation (Brillas and Martinez-Huitle, 2015; He et al., 2021).

The driving forces for the adsorption of organic pollutants by electro-coagulation generally involve van der Waals force, π-π bond, electrostatic interaction, hydrogen bond, ion or ligand exchange, and hydrophobic effect (Lin et al., 2015). Among them, the van der Waals force and π-π bond are impossible due to the lacking of π electrons and the low polarizabilities of the PFAS molecules (Du et al., 2014). Theoretically, PFAS with a charged head group (e.g., –COO^– and –SO₃^–) may be adsorbed by the hydroxyl groups on the metal oxides surface (e.g., AlOOH and α-Fe₂O₃) through ligand exchange and hydrogen bonding (Gao and Chorover, 2012). Gao et al. (2012) demonstrated that PFOA was more easily adsorbed onto iron oxides than PFOS because PFOA could form inner-sphere Fe-carboxylate complexes by the hydroxyl-based ligand exchange, whereas the PFOS only formed outer-sphere complexes and hydrogen bonding at the hydroxide surface.

Yang et al. (2016) reported that electrostatic adsorption was dominant in removing PFAS by a Fe electrode at 25 mA/cm². This was confirmed by the fact that the removal of PFAS was decreased with the increase of the pKa value. Meanwhile, it was found that the removal of PFOS by a Fe anode (99.6%) was higher than that by an Al anode (72.9 %) since more ferric hydroxide flocs can be formed near the surface. However, in another study using Zn as the anode to remove PFAS, experimental data and advanced characterizations consistently indicated that the treatment mechanism was not attributed to ligand exchange, electrostatic attraction, or hydrogen bonding (Lin et al., 2015). It was observed that the removal of PFAS was improved with the increase of the C-F chain length (7.69 mmol/g for PFOS and 5.74 mmol/g for PFOA). The PFAS with a longer chain is more hydrophobic, indicating that hydrophobicity played an important role in the adsorption of PFAS by the Zn hydroxide flocs. The difference in removal mechanism between Al and Zn electrodes may be because the Al hydroxide flocs are mainly hydrophilic, while the Zn hydroxide flocs are more hydrophobic (Lin et al., 2015). Furthermore, it was identified that the removal of PFOA by the metal anodes followed the following order: Zn >> Al >> Mg and Fe ( Lin et al., 2015), which is contrary to the above results by Yang et al. (2016). Therefore, it can find that the removal mechanisms of PFAS by electro-coagulation technology are complex and controversial, which may be closely related to the surface property of the anode materials.

4.2 Effect of operating conditions on the electro-coagulation performance of PFAS

4.2.1 Electrode material

The removal of PFAS strongly depends on the physicochemical properties of the hydroxide flocs generated from the metallic anode. Fe, Al, Mg, and Zn are the main anode materials used in the electro-coagulation process, and they display different performances and mechanisms in removing PFAS. For example, the removal efficiency of PFOS was 99.6 % and 52.5 % within 50 min when Fe and Al were used as the anodes, respectively (Yang et al., 2016). The relatively higher performance of the Fe anode is related to the fact that only 2 electrons are required to oxidize 1 Fe atom, while the Al atom requires 3 electrons. Meanwhile, the solubility and the oxidation potential of the anode metal material can also affect the PFOS removal. Hence, it is expected to have a higher ferric ions concentration than Al³⁺. Therefore, it is more likely to form complexes with the carboxylate or sulfonate of PFAS (Yang et al., 2016). The Fe, Al, Mg, and Zn anodes were investigated under identical conditions to identify the preferred anode material. The results indicated that the Zn anode was far more effective in adsorbing PFOA than the other three anode materials (96.7 % vs. 3.6 %–11.3 %) (Lin et al., 2015). However, the reasons for the obvious difference among the four anode materials were not clearly stated.

Recently, some studies reported that the cathode materials could also affect the removal of PFAS and energy consumption (Khandegar and Saroha, 2013; Liu et al., 2018b). Liu et al. (2018b) investigated the removal of PFOA with six combined electrodes (anode-cathode), including Fe-Fe, Fe-Zn, Fe-Al, Al-Al, Zn-Zn, and Al-Zn. It was found that the Al-Zn electrode can achieve the highest removal of PFOA (99.3 %) within 20 min at 9 V with the lowest energy consumption of 0.11 kWh/m³. In another study, the Al rod and stainless steel rod were selected as the cathodes to compare the removal of PFOA with the Zn anode. A significant difference between them after 20 min of operation was observed. The PFOA removal by the Zn-stainless steel system could reach 96.2 %, while that by the Zn-Al system was only 4.1 % when Na₂SO₄ was used as the supporting electrolyte (Wang et al., 2016). This is inconsistent with the previous report by Benhadji et al. (2011). Unfortunately, these studies only focused on the difference in apparent performance, and the underlying mechanisms were not identified. Therefore, it is necessary to systematically study the influence of anode and cathode materials on PFAS removal by electro-coagulation technology.

4.2.2 Current density

The removal of PFAS increases with the increase of current density within a certain range because higher current density produces more flocs from the anode. However, once the optimal current density is exceeded, the removal of PFAS may decrease. This deterioration is attributed to the production of excessive O₂ and H₂ by the water electrolysis, which floats the flocs and thus reduces the effective contact between the flocs and PFAS (Fajardo et al., 2015). For instance, the removal of PFOS could be improved from 38.3 % to 99.6 % when the current density was adjusted from 6.25 to 25.0 mA/cm² using a Fe anode. Further increasing the current density to 37.5 mA/cm² resulted in a decline in the PFOS removal (~90.0 %) (Yang et al., 2016). Notably, there is also an exception where the removal of PFAS is monotonically increased with the increase of current density from 2.4 to 80.0 mA/cm², rather than this bell-shaped curvature. This may be attributed to producing a higher HO^• concentration at a higher current density (Kim et al., 2020b), although more data is needed to prove this.

4.2.3 Electrolyte type

The chemical components in an electrolyte can also affect the removal of PFAS through different influencing mechanisms in an electro-coagulation process. Yang et al. (2016) investigated the removal of PFOS by a Fe anode with four different types of electrolytes, i.e., NaCl, Na₂SO₄, NaNO₃, and NaH₂PO₄. The results showed that the removal performance follows the following trend: NaCl (99.6%) > Na ₂SO₄ (93.7%) > NaNO ₃ (63.5%) > NaH ₂PO₄ (45.4%). Studies have shown that the immersion of metallic Fe in water at open circuit conditions would result in the generation of impenetrable passive ferric oxide film that restricts the anodic dissolution of Fe. With the presence of Cl^–, the passivation of the Fe electrode could be delayed due to its catalytic action (Amani-Ghadim et al., 2011). However, high concentrations of Cl^– may lead to secondary pollutants (e.g., Cl₂ and ClO^–). Meanwhile, the NO₃^– could be adsorbed on the surfaces of the Fe anode and the Fe hydroxide precipitates, which effectively reduce the electrochemical dissolution and polymerization of the flocs (Amani-Ghadim et al., 2011; Lacasa et al., 2011). With NaH₂PO₄ in the electrolyte, the generation of OH^– is significantly slowed down due to the slowest pH change of the solution, eventually leading to the delayed formation of flocs (Yang et al., 2016). In addition, the Fe K-edge EXAFS spectra revealed that the Fe flocs from electro-coagulation primarily consisted of edge-sharing FeO₆ octahedra. With the presence of PO₄^3–, the formation of the FeO₆ corner-sharing linkage is likely to be restricted to form the Fe flocs (van Genuchten et al., 2012). The hydroxyl group on the surface of flocs can also exchange ligands with the PO₄^3– to form an inner-surface complex (Golder et al., 2006). Based on these analyses, Na₂SO₄ seems to be the perfect candidate that not only can prevent the generation of toxic byproducts but also maintain high efficiency for the removal of PFAS.

5 Coupled technology

Although the electrochemical treatment systems hold great promise for the modular treatment of PFAS-contaminated water, their effectiveness is restricted by the high energy consumption and low current efficiency when the concentration of PFAS is low (1 ng/L to 1 μg/L). To overcome these problems, new technology is emerging by connecting or combing other systems (e.g., nanofiltration (NF), ion exchange resin (IER), and ultrasound activation) with the electrochemical technologies with improved the overall effectiveness and practical feasibility of the electrochemical systems (Soriano et al., 2017; Liang et al., 2018; Pica et al., 2019; Soriano et al., 2019; Soriano et al., 2020; Li et al., 2021b; Maldonado et al., 2021; Shi et al., 2021; Li et al., 2022a; Liang et al., 2022b; Xie et al., 2022). These coupled processes and their key operation parameters are summarized in Table S1. Here, several representative cases are discussed, divided into two categories: tandem (Fig.7(a)) and integration (Fig.7(b)) arrangements of multiple technologies. The following are several recent innovative designs, along with some preliminary results that, to some extent, provided some indication for the further development of electrochemical technology.

Fig.7 Schematic illustrations of (a) tandem, (b) integration arrangements (the dotted box presents that the two processes may operate simultaneously to remove PFAS), (c) nanofiltration coupled electro-oxidation process (Adapted from (Soriano et al., 2017)), and (d) reactive electrochemical membrane for PFAS removal.

Full size|PPT slide

5.1 Tandem arrangement

The tandem treatment system can be considered integration of pretreatment (e.g., adsorption and filtration) and electro-oxidation degradation. The first selected study was performed to remove PFOA with a tandem coupling of an IER procedure and the electrochemical oxidation on a Ti₄O₇ ceramic 3D electrode (Liang et al., 2018). The concentrated sample from the IER had a PFOA concentration of 15.5 mg/L, PFOS concentration of 25.5 mg/L, Cl^– concentration of 853 mg/L, and a high total organic carbon (TOC) background of 9880 mg/L. Results showed that the highly concentrated PFOA and PFOS could be reduced to levels below the limits of quantification after 16 h of electro-oxidation treatment. Meanwhile, the estimated energy consumption for treating PFOA and PFOS concentrations to nondetectable levels was 450.0 kWh/m³. This study also indicated that the removal of PFOA (77.2 %) and PFOS (96.5 %) would decrease when the waste contains a higher concentration of TOC. Therefore, it was suggested that PFAS removal might be significantly improved if the TOC concentration can be reduced before the electro-oxidation treatment (Liang et al., 2018). Recently, the author demonstrated a pilot-scale field facility to further evaluate the remediation of PFAS from groundwater by the coupled system (Liang et al., 2022b). The concentrated waste stream from the IXR was treated on-site by electro-oxidation, achieving 80 %–98 % destruction of PFOA and PFOS. Results showed that the treated effluent from the IXR system had undetectable levels of PFOA and PFOS. Furthermore, the energy consumption was orders of magnitude lower than the reported values from stand-alone electrochemical treatment. Hence, the coupling of regenerable IER with electro-oxidation is a promising candidate for high-performance PFAS treatment from groundwater.

To remove the perfluorohexanoic acid (PFHxA) from real industrial processing water, Soriano et al. (2017) proposed a new sequential treatment system consisting of NF separation followed by electro-oxidation degradation of the NF concentrate. The NF membrane first rejected 99.4 % PFHxA at the operating pressure of 20 bar. The retentate with a PFHxA concentration of 870 mg/L was treated in a commercial undivided electrochemical cell equipped with two integration flow-by compartments separated by a bipolar BDD electrode (Fig.7(c)). Results showed that the degradation of PFHxA could reach 98 % and was accompanied by over 95 % mineralization at a current density of 50 A/m² (Soriano et al., 2017). Meanwhile, the energy consumption for the electro-oxidation degradation of PFHxA was only 15.2 kWh/m³. Similarly, another study demonstrated that the NF90 membrane could separate 99.5 % of GenX to achieve a concentrated solution of 4.98 mg/L (Pica et al., 2019). This concentrated GenX was subsequently degraded by the electro-oxidation treatment. The energy and electrode material costs were reduced by more than 1 order of magnitude compared to direct degradation of the raw water (1 mg/L for GenX). These positive results suggested that the sequentially coupled NF and electrochemical oxidation tandem arrangement is an efficient and cost-effective strategy for eliminating PFAS from contaminated water.

5.2 Integration arrangement

The integration treatment approaches were employed to treat PFAS by combining two remediation technologies simultaneously. During this couped process, the shortcomings of each technology can be mutually compensated. Thus, the defluorination and mineralization of PFAS are optimized.

Membrane separation has been proved effective in removing PFAS. However, the process is generally hindered by membrane fouling (Nunes, 2020). Electrochemical oxidation combines the effects of direct electron transfer on the anode and the oxidation by HO^• produced by anodic oxidation of water (Shi et al., 2019). The reactive electrochemical membrane (REM) combines membrane separation and electrochemical oxidation (Fig.7(d)). In such a configuration, the membrane antifouling can be achieved by in situ electro-oxidation (Liu et al., 2020). Also important, when dealing with the low concentration (sub-μg/L) of PFAS, the enriched PFAS by the membrane filtration creates a possibility for efficient electrochemical oxidation at reduced energy consumption with improved current efficiency. For example, an energy-efficient Magnéli phase Ti₄O₇ REM was used to enrich and oxidize PFOA and PFOS (Le et al., 2019). Results indicated that approximately 5-log removal of these compounds was achieved (from 4.14 mg/L to < 86 ng/L for PFOA, from 5 mg/L to 35 ng/L for PFOS) in a single-pass through the REM with a residence time of ~11 s at 3.3 V for PFOA and 3.6 V for PFOS, respectively. Meanwhile, the energy consumptions for the removal of PFOA and PFOS were calculated to be 5.1 kWh/m ³ and 6.7 kWh/m³, which were the lowest value reported for the electrochemical oxidation of PFAS (Le et al., 2019). Recently, the use of a similar combined configuration to effectively remove PFAS has been widely reported (Shi et al., 2019; Lin et al., 2021; Khalid et al., 2022; Yang et al., 2022), and these processes verified the superior performance of REM in comparison to the electro-oxidation process alone without chemical filtration and enrichment.

Earlier, Zhao et al. (2013) reported the degradation of PFOA by electro-oxidation with a SnO₂-Sb/carbon aerogel anode. They found that the degradation was significantly increased with the assistance of ultrasound. Nevertheless, its mechanism is still not clear. Based on this, Xu et al. (2020) further evaluated the degradation of PFOA in a similar system (Ti/SnO₂-Sb/Ce-PbO₂ as the anode) to highlight the synergic effects of electro-oxidation and ultrasound activation. The performance in the defluorination (87.9 %) and the mineralization (86.1 %) of PFOA in the ultrasonic-assisted electro-oxidation system were higher than those without ultrasound (75.9 % and 71.9 %). Such enhanced PFOA degradation with ultrasound assistance was mainly due to the collapse of bubbles, higher HO^• generation (1.82 times), and a faster anodic electron transfer rate. At present, although the understanding of the ultrasound effects is limited, satisfactory performance for PFAS removal was achieved from these emerging strategies. Overall, the coupling processes still need to be explored in many aspects (e.g., the effect of co-contaminants, the controllability of coupled system, and the design of the modular device) before they can be further improved and become a viable solution for the full-scale treatment of PFAS.

6 Conclusions and outlook

6.1 Conclusions

This review provides a comprehensive summary of the current state of knowledge on the PFAS treatment using electrochemical processes, including electro-oxidation, electro-adsorption, electro-coagulation, and coupled technologies. To better compare different electrochemical PFAS removal technologies, the removal mechanism, energy consumption, and key operating parameters of these methods are summarized in Table S2. Based on the discussions and analysis of the published literature, several conclusions could be drawn, including:

1) Electro-oxidation system usually requires a high applied potential of > 3.0 V ( vs. SHE) to induce electron transfer from PFAS to the anode, which is the main reaction pathway for its degradation. Therefore, electrodes must be anodically stable. Nevertheless, very few low-cost anodic materials can maintain high electrical activity over a long period. It remains a challenge to find the electrode material not suffering from dissolution and disintegration at high potential, especially if it is made of toxic metals or nanomaterials (e.g., PbO₂ and CNT).

2) The three electrochemical technologies are suitable for different application scenarios due to their unique characteristics. For example, many water environments (e.g., groundwater) containing PFAS have low conductivity, and the addition of supporting electrolytes can cause secondary pollution. This may limit the use of electro-oxidation and electro-coagulation technologies. In addition, with a variety of negatively charged anions and metal ions in the wastewater, the electro-adsorption technology will not be efficient due to the competitive effect. Table S3 presents the advantages and disadvantages of three electrochemical technologies in removing PFAS. Moreover, the electrochemical system is modular, scalable, autonomous, and robust. Thus, electrochemical cells as a decentralized point-of-use and point-of-entry water treatment technology seem to be the ideal application scenario.

3) At present, the reported electrochemical strategies mainly focus on removing PFOA and PFOS, whereas the treatment of short-chain PFAS is rarely reported. Furthermore, the reported PFAS removal performance may be overestimated, as most studies were conducted in synthetic wastewater where PFAS concentrations (mg/L) were much higher than in actual contaminated water (ng/L–μg/L).

4) The development of emerging electrochemical processes strongly relies on innovations in fundamental material science. Most of the papers published to date focus on the design of anode materials to improve PFAS removal. However, studies on the remediation mechanisms and influence factors of electrochemical technologies are often superficial, rather than systematic and comprehensive, and even contradictory.

5) According to the literature reported in recent years, it can be found that the coupled technology (e.g., IER/electro-oxidation and REM systems) has received considerable attention and is expected to be an effective method for PFAS treatment due to its low cost and highly efficient performance.

6) In these reported strategies, the tested initial concentration range of PFAS and the water environments have significant variations, making it difficult to compare these electrochemical processes objectively. Although some preliminary studies also compared and analyzed the future development prospects of these technologies in terms of costs and energy consumption, such considerations may be premature for these technologies due to their rapid development.

6.2 Outlook

Despite all the remarkable achievements made in the past years in the remediation of PFAS in water through electrochemical technologies, there are still some investigations that need to be addressed (Table S4):

1) The limitation of electrochemical technologies for practical application is that the electrode surface is highly reactive toward both target PFAS molecules and various water matrix components (e.g., NOM and Cl^–). These components are usually much more abundant than trace amounts of PFAS. Consequently, the fraction of electron or active sites on the electrode surface for the PFAS removal is limited and thus results in very low efficacy with high energy consumption and high operation cost. Therefore, novel strategies to minimize the competition effects of the coexisting species are highly demanded. For example, nanocomposite should be designed to achieve superior selectivity and precise recognition of target compounds in a complex water matrix, which can preferentially interact with PFAS to alleviate the selectivity and competition problems.

2) Complete mineralization of PFAS by electro-oxidation is currently still difficult. Any residual intermediates might be more toxic or harmful than the original molecules. Hence, the characterization of the decomposition intermediates and analysis of their toxicities should also be focused on. The influences of operational conditions and water chemistry parameters on the formation of these intermediates are worth to be further investigated. In addition, attention should also be paid to the generation of fluoride ions in water, the subsequent treatment of which has rarely been reported until now.

3) More efforts should be dedicated to the technology assessment under more realistic environments to deliver practical engineering applications. For example, the initial concentration of PFAS should be comparable to the concentration range of actual contaminated water. Attention should also be paid to short-chain compounds other than PFOA and PFOS. Longer-term experiments and on-site demonstration tests should be carried out, allowing a more accurate estimation of the impacts of interfering components on the electrode used (e.g., scaling or corrosion).

4) Coupled technology as a novel treatment train should be continuously developed to improve the overall effectiveness and practical feasibility, as well as to maximize the economic and environmental benefits of the electrochemical systems. Besides, the design and development of coupled devices are also highly needed, although it will be challenging.

5) Electrochemical technology with excellent selectivity for PFAS needs to be developed. Electroialydsis has been widely used in wastewater treatment, especially in the selective removal of toxic heavy metal ions. It is significant to explore the selective removal of PFAS by electrodialysis in future research.

6) Developing low-cost and energy-efficient electrochemical technologies is crucial to remedy PFAS pollution. Meanwhile, it is necessary to carry out the techno-economic assessment (TEA) and life-cycle analysis (LCA) of electrochemical technologies and compare them with the existing PFAS treatment schemes. With these assessments, residual disposal and management should also be considered.

Overall, significant improvements have been made to the PFAS treatment by electrochemical technologies. With the explorations of novel electrodes and process designs for the PFAS remediation in water, researchers must carefully, completely, and fairly evaluate the electrode suitability and the treatment results to avoid false conclusions which might result in unnecessary academic misleading. In addition, more efforts are encouraged to scrutinize various controversial results in this field (Table S4), such as the effect of coexisting ions and the exact decomposition mechanisms of electro-oxidation of PFAS, which will be more useful for the further development of related technologies.

Acknowledgements

This work was supported by the National Natural Science Foundation of China (No. 52170068 and U21A20161) and the Open Project of State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology (No. QAK202108).

Electronic Supplementary Material

Supplementary material is available in the online version of this article at https://doi.org/10.1007/s11783-023-1618-z and is accessible for authorized users.

References

1 Amani-Ghadim A R , Aber S , Olad A , Ashassi-Sorkhabi H . (2011). Influence of anions on Reactive Red 43 removal in electrochemical coagulation process. Electrochimica Acta, 56( 3): 1373– 1380

2 Azizi O , Hubler D , Schrader G , Farrell J , Chaplin B P . (2011). Mechanism of perchlorate formation on boron-doped diamond film anodes. Environmental Science & Technology, 45( 24): 10582– 10590

PMID

3 Bao J , Yu W J , Liu Y , Wang X , Liu Z Q , Duan Y F . (2020). Removal of perfluoroalkanesulfonic acids (PFSAs) from synthetic and natural groundwater by electrocoagulation. Chemosphere, 248 : 125951

PMID

4 Barisci S , Suri R . (2020). Electrooxidation of short and long chain perfluorocarboxylic acids using boron doped diamond electrodes. Chemosphere, 243 : 125349

PMID

5 Barisci S , Suri R . (2021). Electrooxidation of short- and long-chain perfluoroalkyl substances (PFASs) under different process conditions. Journal of Environmental Chemical Engineering, 9( 4): 105323

6 Bayram E , Ayranci E . (2010). Electrochemically enhanced removal of polycyclic aromatic basic dyes from dilute aqueous solutions by activated carbon cloth electrodes. Environmental Science & Technology, 44( 16): 6331– 6336

PMID

7 Bejan D , Guinea E , Bunce N J . (2012). On the nature of the hydroxyl radicals produced at boron-doped diamond and Ebonex (R) anodes. Electrochimica Acta, 69 : 275– 281

8 Benhadji A, Taleb Ahmed M, Maachi R ( 2011). Electrocoagulation and effect of cathode materials on the removal of pollutants from tannery wastewater of Rouiba. Desalination, 277( 1– 3): 128– 134

9 Bentel M J , Yu Y , Xu L , Li Z , Wong B M , Men Y , Liu J . (2019). Defluorination of per- and polyfluoroalkyl substances (PFASs) with hydrated electrons: Structural dependence and implications to PFAS remediation and management. Environmental Science & Technology, 53( 7): 3718– 3728

PMID

10 Brillas E, Martinez-Huitle C A ( 2015). Decontamination of wastewaters containing synthetic organic dyes by electrochemical methods. An updated review. Applied Catalysis B: Environmental, 166– 167: 603– 643

11 Carter K E , Farrell J . (2008). Oxidative destruction of perfluorooctane sulfonate using boron-doped diamond film electrodes. Environmental Science & Technology, 42( 16): 6111– 6115

PMID

12 Chaplin B P . (2014). Critical review of electrochemical advanced oxidation processes for water treatment applications. Environmental Science. Processes & Impacts, 16( 6): 1182– 1203

PMID

13 Chaplin B P . (2019). The prospect of electrochemical technologies advancing worldwide water treatment. Accounts of Chemical Research, 52( 3): 596– 604

PMID

14 Du Z , Deng S , Bei Y , Huang Q , Wang B , Huang J , Yu G . (2014). Adsorption behavior and mechanism of perfluorinated compounds on various adsorbents-a review. Journal of Hazardous Materials, 274 : 443– 454

PMID

15 Duan X , Wang W , Wang Q , Sui X , Li N , Chang L . (2020). Electrocatalytic degradation of perfluoroocatane sulfonate (PFOS) on a 3D graphene-lead dioxide (3DG-PbO₂) composite anode: electrode characterization, degradation mechanism and toxicity. Chemosphere, 260 : 127587

PMID

16 Duinslaeger N , Radjenovic J . (2022). Electrochemical degradation of per- and polyfluoroalkyl substances (PFAS) using low-cost graphene sponge electrodes. Water Research, 213 : 118148

PMID

17 Fajardo A S , Rodrigues R F , Martins R C , Castro L M , Quinta-Ferreira R M . (2015). Phenolic wastewaters treatment by electrocoagulation process using Zn anode. Chemical Engineering Journal, 275 : 331– 341

18 Gao X D , Chorover J . (2012). Adsorption of perfluorooctanoic acid and perfluorooctanesulfonic acid to iron oxide surfaces as studied by flow-through ATR-FTIR spectroscopy. Environmental Chemistry, 9( 2): 148– 157

19 Geng P , Su J Y , Miles C , Comninellis C , Chen G H . (2015). Highly-ordered magneli Ti₄O₇ nanotube arrays as effective anodic material for electro-oxidation. Electrochimica Acta, 153 : 316– 324

20 Giesy J P , Kannan K . (2001). Global distribution of perfluorooctane sulfonate in wildlife. Environmental Science & Technology, 35( 7): 1339– 1342

PMID

21 Giesy J P , Kannan K . (2002). Perfluorochemical surfactants in the environment. Environmental Science & Technology, 36( 7): 146A– 152A

PMID

22 Golder A K , Samanta A N , Ray S . (2006). Removal of phosphate from aqueous solutions using calcined metal hydroxides sludge waste generated from electrocoagulation. Separation and Purification Technology, 52( 1): 102– 109

23 Grandjean P . (2018). Delayed discovery, dissemination, and decisions on intervention in environmental health: a case study on immunotoxicity of perfluorinated alkylate substances. Environmental Health, 17( 1): 62

PMID

24 Haro M , Rasines G , Macias C , Ania C O . (2011). Stability of a carbon gel electrode when used for the electro-assisted removal of ions from brackish water. Carbon, 49( 12): 3723– 3730

25 He H , Yu Q J , Lai C C , Zhang C , Liu M H , Huang B , Pu H P , Pan X J . (2021). The treatment of black-odorous water using tower bipolar electro-flocculation including the removal of phosphorus, turbidity, sulfion, and oxygen enrichment. Frontiers of Environmental Science & Engineering, 15( 2): 18

26 Hogue C . (2019). Governments endorse global PFOA ban, with some exemptions. Chemical and Engineering News, 97( 19): 5– 5

27 Houde M , Martin J W , Letcher R J , Solomon K R , Muir D C G . (2006). Biological monitoring of polyfluoroalkyl substances: a review. Environmental Science & Technology, 40( 11): 3463– 3473

PMID

28 Huang D , Wang K , Niu J , Chu C , Weon S , Zhu Q , Lu J , Stavitski E , Kim J H . (2020). Amorphous Pd-loaded Ti₄O₇ electrode for direct anodic destruction of perfluorooctanoic acid. Environmental Science & Technology, 54( 17): 10954– 10963

PMID

29 Huang S Y , Fan C S , Hou C H . (2014). Electro-enhanced removal of copper ions from aqueous solutions by capacitive deionization. Journal of Hazardous Materials, 278 : 8– 15

PMID

30 Hwang J H , Li Sip Y Y , Kim K T , Han G , Rodriguez K L , Fox D W , Afrin S , Burnstine-Townley A , Zhai L , Lee W H . (2022). Nanoparticle-embedded hydrogel synthesized electrodes for electrochemical oxidation of perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS). Chemosphere, 296 : 134001

PMID

31 Kelly B C , Ikonomou M G , Blair J D , Surridge B , Hoover D , Grace R , Gobas F A . (2009). Perfluoroalkyl contaminants in an Arctic marine food web: trophic magnification and wildlife exposure. Environmental Science & Technology, 43( 11): 4037– 4043

PMID

32 Key B D , Howell R D , Criddle C S . (1997). Fluorinated organics in the biosphere. Environmental Science & Technology, 31( 9): 2445– 2454

33 Khalid Y S , Misal S N , Mehraeen S , Chaplin B P . (2022). Reactive-transport modeling of electrochemical oxidation of perfluoroalkyl substances in porous flow-through electrodes. ACS ES&T Engineering, 2( 4): 713– 725

34 Khandegar V , Saroha A K . (2013). Electrocoagulation for the treatment of textile industry effluent-a review. Journal of Environmental Management, 128 : 949– 963

PMID

35 Kim K , Baldaguez Medina P , Elbert J , Kayiwa E , Cusick R D , Men Y J , Su X . (2020a). Molecular tuning of redox-copolymers for selective electrochemical remediation. Advanced Functional Materials, 30( 52): 2004635

36 Kim M K , Kim T , Kim T K , Joo S W , Zoh K D . (2020b). Degradation mechanism of perfluorooctanoic acid (PFOA) during electrocoagulation using Fe electrode. Separation and Purification Technology, 247 : 116911

37 Lacasa E , Canizares P , Saez C , Fernandez F J , Rodrigo M A . (2011). Removal of nitrates from groundwater by electrocoagulation. Chemical Engineering Journal, 171( 3): 1012– 1017

38 Le T X H , Haflich H , Shah A D , Chaplin B P . (2019). Energy-efficient electrochemical oxidation of perfluoroalkyl substances using a Ti₄O₇ reactive electrochemical membrane anode. Environmental Science & Technology Letters, 6( 8): 504– 510

39 Li L , Wang Y Y , Huang Q G . (2021a). First-principles study of the degradation of perfluorooctanesulfonate and perfluorobutanesulfonate on a Magneli phase Ti₄O₇ anode. ACS ES&T Water, 1( 8): 1737– 1744

40 Li M , Jin Y T , Yan J F , Liu Z , Feng N X , Han W , Huang L W , Li Q K , Yeung K L , Zhou S Q , Mo C H . (2022a). Exploration of perfluorooctane sulfonate degradation properties and mechanism via electron-transfer dominated radical process. Water Research, 215 : 118259

PMID

41 Li M , Mo C H , Luo X , He K Y , Yan J F , Wu Q , Yu P F , Han W , Feng N X , Yeung K L , Zhou S Q . (2021b). Exploring key reaction sites and deep degradation mechanism of perfluorooctane sulfonate via peroxymonosulfate activation under electrocoagulation process. Water Research, 207 : 117849

PMID

42 Li X , Chen S , Quan X , Zhang Y . (2011). Enhanced adsorption of PFOA and PFOS on multiwalled carbon nanotubes under electrochemical assistance. Environmental Science & Technology, 45( 19): 8498– 8505

PMID

43 Li Y F, Hu C Y, Lee Y C, Lo S L ( 2022b). Effects of zinc salt addition on perfluorooctanoic acid (PFOA) removal by electrocoagulation with aluminum electrodes. Chemosphere, 288( Pt 3): 132665

PMID

44 Liang Q J , Gao Y , Li Z G , Cai J Y , Chu N , Hao W , Jiang Y , Zeng R J X . (2022a). Electricity-driven ammonia oxidation and acetate production in microbial electrosynthesis systems. Frontiers of Environmental Science & Engineering, 16( 4): 42

45 Liang S T , Mora R , Huang Q G , Casson R , Wang Y Y , Woodard S , Anderson H . (2022b). Field demonstration of coupling ion-exchange resin with electrochemical oxidation for enhanced treatment of per- and polyfluoroalkyl substances (PFAS) in groundwater. Chemical Engineering Journal Advances, 9 : 100216

46 Liang S T , Pierce R Jr , Lin H , Chiang S Y , Huang Q . (2018). Electrochemical oxidation of PFOA and PFOS in concentrated waste streams. Remediation Journal, 28( 2): 127– 134

47 Lin H , Niu J , Ding S , Zhang L . (2012). Electrochemical degradation of perfluorooctanoic acid (PFOA) by Ti/SnO₂-Sb, Ti/SnO₂-Sb/PbO₂ and Ti/SnO₂-Sb/MnO₂ anodes. Water Research, 46( 7): 2281– 2289

PMID

48 Lin H , Niu J F , Liang S T , Wang C , Wang Y J , Jin F Y , Luo Q , Huang Q G . (2018). Development of macroporous Magneli phase Ti₄O₇ ceramic materials: As an efficient anode for mineralization of poly- and perfluoroalkyl substances. Chemical Engineering Journal, 354 : 1058– 1067

49 Lin H , Niu J F , Xu J L , Li Y , Pan Y H . (2013). Electrochemical mineralization of sulfamethoxazole by Ti/SnO₂-Sb/Ce-PbO₂ anode: Kinetics, reaction pathways, and energy cost evolution. Electrochimica Acta, 97 : 167– 174

50 Lin H , Wang Y , Niu J , Yue Z , Huang Q . (2015). Efficient sorption and removal of perfluoroalkyl acids (PFAAs) from aqueous solution by metal hydroxides generated in situ by electrocoagulation. Environmental Science & Technology, 49( 17): 10562– 10569

PMID

51 Lin H , Xiao R , Xie R , Yang L , Tang C , Wang R , Chen J , Lv S , Huang Q . (2021). Defect engineering on a Ti₄O₇ electrode by Ce³⁺ doping for the efficient electrooxidation of perfluorooctanesulfonate. Environmental Science & Technology, 55( 4): 2597– 2607

PMID

52 Lin M H , Mehraeen S , Cheng G , Rusinek C , Chaplin B P . (2020). Role of near-electrode solution chemistry on bacteria attachment and poration at low applied potentials. Environmental Science & Technology, 54( 1): 446– 455

PMID

53 Lindstrom A B , Strynar M J , Libelo E L . (2011). Polyfluorinated compounds: past, present, and future. Environmental Science & Technology, 45( 19): 7954– 7961

PMID

54 Liu G S , Zhou H , Teng J , You S J . (2019a). Electrochemical degradation of perfluorooctanoic acid by macro- porous titanium suboxide anode in the presence of sulfate. Chemical Engineering Journal, 371 : 7– 14

55 Liu L , Li D , Li C , Ji R , Tian X . (2018a). Metal nanoparticles by doping carbon nanotubes improved the sorption of perfluorooctanoic acid. Journal of Hazardous Materials, 351 : 206– 214

PMID

56 Liu L , Liu Y , Che N , Gao B , Li C . (2021a). Electrochemical adsorption of perfluorooctanoic acid on a novel reduced graphene oxide aerogel loaded with Cu nanoparticles and fluorine. Journal of Hazardous Materials, 416 : 125866

PMID

57 Liu L Q , Qu Y X , Huang J , Weber R . (2021b). Per- and polyfluoroalkyl substances (PFASs) in Chinese drinking water: Risk assessment and geographical distribution. Environmental Sciences Europe, 33( 1): 6

58 Liu Y , Chen S , Quan X , Yu H , Zhao H , Zhang Y . (2015). Efficient mineralization of perfluorooctanoate by electro-Fenton with H₂O₂ electro-generated on hierarchically porous carbon. Environmental Science & Technology, 49( 22): 13528– 13533

PMID

59 Liu Y , Fan X , Quan X , Fan Y , Chen S , Zhao X . (2019b). Enhanced perfluorooctanoic acid degradation by electrochemical activation of sulfate solution on B/N codoped diamond. Environmental Science & Technology, 53( 9): 5195– 5201

PMID

60 Liu Y , Gao G , Vecitis C D . (2020). Prospects of an electroactive carbon nanotube membrane toward environmental applications. Accounts of Chemical Research, 53( 12): 2892– 2902

PMID

61 Liu Y , Hu X M , Zhao Y , Wang J , Lu M X , Peng F H , Bao J . (2018b). Removal of perfluorooctanoic acid in simulated and natural waters with different electrode materials by electrocoagulation. Chemosphere, 201 : 303– 309

PMID

62 Lu D , Sha S , Luo J , Huang Z , Zhang Jackie X . (2020). Treatment train approaches for the remediation of per- and polyfluoroalkyl substances (PFAS): a critical review. Journal of Hazardous Materials, 386 : 121963

PMID

63 Luo G M , Wang Y Z , Gao L X , Zhang D Q , Lin T . (2018). Graphene bonded carbon nanofiber aerogels with high capacitive deionization capability. Electrochimica Acta, 260 : 656– 663

64 Ma C Y , Huang S C , Chou P H , Den W , Hou C H . (2016). Application of a multiwalled carbon nanotube-chitosan composite as an electrode in the electrosorption process for water purification. Chemosphere, 146 : 113– 120

PMID

65 Maldonado V Y , Becker M F , Nickelsen M G , Witt S E . (2021). Laboratory and semi-pilot scale study on the electrochemical treatment of perfluoroalkyl acids from ion exchange still bottoms. Water, 13( 20): 2873

66 Martínez-Huitle C A , Rodrigo M A , Sirés I , Scialdone O . (2015). Single and coupled electrochemical processes and reactors for the abatement of organic water pollutants: a critical review. Chemical Reviews, 115( 24): 13362– 13407

PMID

67 Baldaguez Medina P , Cotty S , Kim K , Elbert J , Su X . (2021). Emerging investigator series: electrochemically-mediated remediation of GenX using redox-copolymers. Environmental Science: Water Research & Technology, 7( 12): 2231– 2240

68 Mollah MYA, Morkovsky P, Gomes JAG, Kesmez M, Parga J, Cocke DL ( 2004). Fundamentals, present and future perspectives of electrocoagulation. Journal of Hazardous Materials, 114( 1– 3): 199– 210

PMID

69 Mu T , Park M , Kim K Y . (2021). Energy-efficient removal of PFOA and PFOS in water using electrocoagulation with an air-cathode. Chemosphere, 281 : 130956

PMID

70 Nienhauser A B , Ersan M S , Lin Z H , Perreault F , Westerhoff P , Garcia-Segura S . (2022). Boron-doped diamond electrodes degrade short- and long-chain per- and polyfluorinated alkyl substances in real industrial wastewaters. Journal of Environmental Chemical Engineering, 10( 2): 107192

71 Niu J , Li Y , Shang E , Xu Z , Liu J . (2016). Electrochemical oxidation of perfluorinated compounds in water. Chemosphere, 146 : 526– 538

PMID

72 Niu J , Lin H , Gong C , Sun X . (2013). Theoretical and experimental insights into the electrochemical mineralization mechanism of perfluorooctanoic acid. Environmental Science & Technology, 47( 24): 14341– 14349

PMID

73 Niu J , Lin H , Xu J , Wu H , Li Y . (2012). Electrochemical mineralization of perfluorocarboxylic acids (PFCAs) by Ce-doped modified porous nanocrystalline PbO₂ film electrode. Environmental Science & Technology, 46( 18): 10191– 10198

PMID

74 Niu Z J , Wang Y J , Lin H , Jin F Y , Li Y , Niu J F . (2017). Electrochemically enhanced removal of perfluorinated compounds (PFCs) from aqueous solution by CNTs-graphene composite electrode. Chemical Engineering Journal, 328 : 228– 235

75 Nunes S P ( 2020). Can fouling in membranes be ever defeated? Current Opinion in Chemical Engineering, 28: 90– 95

76 Nzeribe N , Crimi M , Mededovic Thagard S , Holsen T M . (2019). Physico-chemical processes for the treatment of per- and polyfluoroalkyl substances (PFAS): a review blossom. Critical Reviews in Environmental Science and Technology, 49( 10): 866– 915

77 Opoku-Duah S , Johnson D . (2020). Removal of perfluorooctanoic acid and microcystins from drinking water by electrocoagulation. Journal of Chemistry, 2020 : 1836264

78 Pica N E , Funkhouser J , Yin Y , Zhang Z , Ceres D M , Tong T , Blotevogel J . (2019). Electrochemical oxidation of hexafluoropropylene oxide dimer acid (GenX): mechanistic insights and efficient treatment train with nanofiltration. Environmental Science & Technology, 53( 21): 12602– 12609

PMID

79 Pierpaoli M , Szopińska M , Wilk B K , Sobaszek M , Łuczkiewicz A , Bogdanowicz R , Fudala-Książek S . (2021). Electrochemical oxidation of PFOA and PFOS in landfill leachates at low and highly boron-doped diamond electrodes. Journal of Hazardous Materials, 403 : 123606

PMID

80 Qi Z L , You S J , Liu R B , Chuah C J . (2020). Performance and mechanistic study on electrocoagulation process for municipal wastewater treatment based on horizontal bipolar electrodes. Frontiers of Environmental Science & Engineering, 14( 3): 40

81 Quan Q , Wen H , Han S , Wang Z , Shao Z , Chen M . (2020). Fluorous-core nanoparticle-embedded hydrogel synthesized via tandem photo-controlled radical polymerization: facilitating the separation of perfluorinated alkyl substances from water. ACS Applied Materials & Interfaces, 12( 21): 24319– 24327

PMID

82 Radjenovic J, Duinslaeger N, Avval S S, Chaplin B P ( 2020). Facing the challenge of poly- and perfluoroalkyl substances in water: Is electrochemical oxidation the answer? Environmental Science & Technology, 54( 23): 14815– 14829

PMID

83 Rahman M F , Peldszus S , Anderson W B . (2014). Behaviour and fate of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in drinking water treatment: a review. Water Research, 50 : 318– 340

PMID

84 Ren Y F , Zheng W T , Duan X G , Goswami N , Liu Y B . (2022). Recent advances in electrochemical removal and recovery of phosphorus from water: a review. Environmental Functional Materials, 1( 1): 10– 20

85 Saeidi N , Kopinke F D , Georgi A . (2021). Controlling adsorption of perfluoroalkyl acids on activated carbon felt by means of electrical potentials. Chemical Engineering Journal, 416 : 129070

86 Santos D , Lopes A , Pacheco M J , Gomes A , Ciriaco L . (2014). The oxygen evolution reaction at Sn-Sb oxide anodes: influence of the oxide preparation mode. Journal of the Electrochemical Society, 161( 9): H564– H572

87 Sharma S , Shetti N P , Basu S , Nadagouda M N , Aminabhavi T M . (2022). Remediation of per- and polyfluoroalkyls (PFAS) via electrochemical methods. Chemical Engineering Journal, 430 : 132895

88 Shen Z C , Zhan L , Xu Z M . (2022). Thermal defluorination behaviors of PFOS, PFOA and PFBS during regeneration of activated carbon by molten salt. Frontiers of Environmental Science & Engineering, 16( 8): 103

89 Shi H , Chiang S D , Wang Y , Wang Y , Liang S , Zhou J , Fontanez R , Gao S , Huang Q . (2021). An electrocoagulation and electrooxidation treatment train to remove and degrade per- and polyfluoroalkyl substances in aqueous solution. Science of the Total Environment, 788 : 147723

PMID

90 Shi H , Wang Y , Li C , Pierce R , Gao S , Huang Q . (2019). Degradation of perfluorooctanesulfonate by reactive electrochemical membrane composed of Magnéli phase titanium suboxide. Environmental Science & Technology, 53( 24): 14528– 14537

PMID

91 Shrestha B , Ezazi M , Ajayan S , Kwon G . (2021). Reversible adsorption and desorption of PFAS on inexpensive graphite adsorbents via alternating electric field. RSC Advances, 11( 55): 34652– 34659

PMID

92 Si Y X , Zhang F , Hong C , Li G H , Zhang H C , Liu D . (2021). Effect of current density on groundwater arsenite removal performance using air cathode electrocoagulation. Frontiers of Environmental Science & Engineering, 15( 6): 112

93 Song S , Fan J Q , He Z Q , Zhan L Y , Liu Z W , Chen J M , Xu X H . (2010). Electrochemical degradation of azo dye CI reactive red 195 by anodic oxidation on Ti/SnO₂-Sb/PbO₂ electrodes. Electrochimica Acta, 55( 11): 3606– 3613

94 Soriano Á , Gorri D , Urtiaga A . (2017). Efficient treatment of perfluorohexanoic acid by nanofiltration followed by electrochemical degradation of the NF concentrate. Water Research, 112 : 147– 156

PMID

95 Soriano A , Gorri D , Urtiaga A . (2019). Membrane preconcentration as an efficient tool to reduce the energy consumption of perfluorohexanoic acid electrochemical treatment. Separation and Purification Technology, 208 : 160– 168

96 Soriano A , Schaefer C , Urtiaga A . (2020). Enhanced treatment of perfluoroalkyl acids in groundwater by membrane separation and electrochemical oxidation. Chemical Engineering Journal Advances, 4 : 100042

97 Steenland K , Fletcher T , Savitz D A . (2010). Epidemiologic evidence on the health effects of perfluorooctanoic acid (PFOA). Environmental Health Perspectives, 118( 8): 1100– 1108

PMID

98 Taves D R . (1968). Evidence that there are two forms of fluoride in human serum. Nature, 217( 5133): 1050– 1051

PMID

99 Tian Y H , Xing J Y , Huyan C X , Zhu C Z , Du D , Zhu W L , Lin Y H , Chowdhury I . (2021). Electrically controlled anion exchange based on a polypyrrole/carbon cloth composite for the removal of perfluorooctanoic acid. ACS ES&T Water, 1( 12): 2504– 2512

100 Trautmann A M , Schell H , Schmidt K R , Mangold K M , Tiehm A . (2015). Electrochemical degradation of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in groundwater. Water Science and Technology, 71( 10): 1569– 1575

PMID

101 Trojanowicz M , Bojanowska-Czajka A , Bartosiewicz I , Kulisa K . (2018). Advanced oxidation/reduction processes treatment for aqueous perfluorooctanoate (PFOA) and perfluorooctanesulfonate (PFOS): a review of recent advances. Chemical Engineering Journal, 336 : 170– 199

102 Uwayezu J N , Carabante I , Lejon T , van Hees P , Karlsson P , Hollman P , Kumpiene J . (2021). Electrochemical degradation of per- and poly-fluoroalkyl substances using boron-doped diamond electrodes. Journal of Environmental Management, 290 : 112573

PMID

103 van Genuchten C M , Addy S E A , Peña J , Gadgil A J . (2012). Removing arsenic from synthetic groundwater with iron electrocoagulation: an Fe and As K-edge EXAFS study. Environmental Science & Technology, 46( 2): 986– 994

PMID

104 Veciana M , Bräunig J , Farhat A , Pype M L , Freguia S , Carvalho G , Keller J , Ledezma P . (2022). Electrochemical oxidation processes for PFAS removal from contaminated water and wastewater: fundamentals, gaps and opportunities towards practical implementation. Journal of Hazardous Materials, 434 : 128886

PMID

105 Wang P , Zhang M , Li Q , Lu Y . (2021). Atmospheric diffusion of perfluoroalkyl acids emitted from fluorochemical industry and its associated health risks. Environment International, 146 : 106247

PMID

106 Wang Q N , Liu M Y , Zhao H Y , Chen Y , Xiao F , Chu W H , Zhao G H . (2019a). Efficiently degradation of perfluorooctanoic acid in synergic electrochemical process combining cathodic electro-Fenton and anodic oxidation. Chemical Engineering Journal, 378 : 122071

107 Wang S Y , Li X N , Zhang Y B , Quan X , Chen S , Yu H T , Zhao H M . (2014a). Electrochemically enhanced adsorption of PFOA and PFOS on multiwalled carbon nanotubes in continuous flow mode. Chinese Science Bulletin, 59( 23): 2890– 2897

108 Wang X Y , Xie Y , Yang G Z , Hao J M , Ma J , Ning P . (2020a). Enhancement of the electrocatalytic oxidation of antibiotic wastewater over the conductive black carbon-PbO₂ electrode prepared using novel green approach. Frontiers of Environmental Science & Engineering, 14( 2): 22

109 Wang Y , Lin H , Jin F , Niu J , Zhao J , Bi Y , Li Y . (2016). Electrocoagulation mechanism of perfluorooctanoate (PFOA) on a zinc anode: influence of cathodes and anions. Science of the Total Environment, 557-558 : 542– 550

PMID

110 Wang Y Y , Li L , Wang Y F , Shi H H , Wang L , Huang Q G . (2022). Electrooxidation of perfluorooctanesulfonic acid on porous Magnéli phase titanium suboxide anodes: impact of porous structure and composition. Chemical Engineering Journal, 431 : 133929

111 Wang Y Y , Pierce R , Shi H H , Li C G , Huang Q G . (2020b). Electrochemical degradation of perfluoroalkyl acids by titanium suboxide anodes. Environmental Science: Water Research & Technology, 6( 1): 144– 152

112 Wang Z , Cousins I T , Scheringer M , Buck R C , Hungerbühler K . (2014b). Global emission inventories for C4–C14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, Part I: production and emissions from quantifiable sources. Environment International, 70 : 62– 75

PMID

113 Wu M S , Liao T L , Wang Y Y , Wan C C . (2004). Assessment of the wettability of porous electrodes for lithium-ion batteries. Journal of Applied Electrochemistry, 34( 8): 797– 805

114 Xie J , Zhang C , David Waite T . (2022). Integrated flow anodic oxidation and ultrafiltration system for continuous defluorination of perfluorooctanoic acid (PFOA). Water Research, 216 : 118319

PMID

115 Xu L , Qian X B , Wang K X , Fang C , Niu J F . (2020). Electrochemical mineralization mechanisms of perfluorooctanoic acid in water assisted by low frequency ultrasound. Journal of Cleaner Production, 263 : 121546

116 Xu Z , Yu Y , Liu H , Niu J . (2017). Highly efficient and stable Zr-doped nanocrystalline PbO₂ electrode for mineralization of perfluorooctanoic acid in a sequential treatment system. Science of the Total Environment, 579 : 1600– 1607

PMID

117 Yang B , Han Y N , Yu G , Zhuo Q F , Deng S B , Wu J H , Zhang P X . (2016). Efficient removal of perfluoroalkyl acids (PFAAs) from aqueous solution by electrocoagulation using iron electrode. Chemical Engineering Journal, 303 : 384– 390

118 Yang B , Jiang C , Yu G , Zhuo Q , Deng S , Wu J , Zhang H . (2015). Highly efficient electrochemical degradation of perfluorooctanoic acid (PFOA) by F-doped Ti/SnO₂ electrode. Journal of Hazardous Materials, 299 : 417– 424

PMID

119 Yang L , He L , Xue J , Ma Y , Xie Z , Wu L , Huang M , Zhang Z . (2020). Persulfate-based degradation of perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) in aqueous solution: Review on influences, mechanisms and prospective. Journal of Hazardous Materials, 393 : 122405

PMID

120 Yang L H, Yang W J, Lv S H, Zhu T T, Adeel Sharif H M, Yang C, Du J, Lin H ( 2022). Is HFPO-DA (GenX) a suitable substitute for PFOA? A comprehensive degradation comparison of PFOA and GenX via electrooxidation. Environmental Research, 204(Pt A): 111995

PMID

121 Yang S S , Fernando S , Holsen T M , Yang Y . (2019). Inhibition of perchlorate formation during the electrochemical oxidation of perfluoroalkyl acid in groundwater. Environmental Science & Technology Letters, 6( 12): 775– 780

PMID

122 You S J , Liu B , Gao Y F , Wang Y , Tang C Y Y , Huang Y B , Ren N Q . (2016). Monolithic porous magneli-phase Ti₄O₇ for electro-oxidation treatment of industrial wastewater. Electrochimica Acta, 214 : 326– 335

123 Zhao H Y, Gao J X, Zhao G H, Fan J Q, Wang Y B, Wang Y J ( 2013). Fabrication of novel SnO ₂ -Sb/carbon aerogel electrode for ultrasonic electrochemical oxidation of perfluorooctanoate with high catalytic efficiency. Applied Catalysis B: Environmental, 136– 137: 278– 286

124 Zhong C Q , Wei K J , Han W Q , Wang L J , Sun X Y , Li J S . (2013). Electrochemical degradation of tricyclazole in aqueous solution using Ti/SnO₂-Sb/PbO₂ anode. Journal of Electroanalytical Chemistry, 705 : 68– 74

125 Zhuo Q , Deng S , Yang B , Huang J , Yu G . (2011). Efficient electrochemical oxidation of perfluorooctanoate using a Ti/SnO₂-Sb-Bi anode. Environmental Science & Technology, 45( 7): 2973– 2979

PMID

126 Zhuo Q , Li X , Yan F , Yang B , Deng S , Huang J , Yu G . (2014). Electrochemical oxidation of 1H,1H,2H,2H-perfluorooctane sulfonic acid (6:2 FTS) on DSA electrode: operating parameters and mechanism. Journal of Environmental Sciences (China), 26( 8): 1733– 1739

PMID

127 Zhuo Q F , Deng S B , Yang B , Huang J , Wang B , Zhang T T , Yu G . (2012). Degradation of perfluorinated compounds on a boron-doped diamond electrode. Electrochimica Acta, 77 : 17– 22

128 Zhuo Q F , Luo M Q , Guo Q W , Yu G , Deng S B , Xu Z C , Yang B , Liang X L . (2016). Electrochemical oxidation of environmentally persistent perfluorooctane sulfonate by a novel Lead dioxide anode. Electrochimica Acta, 213 : 358– 367