English

• 1. Introduction

  Sanitary landfills are a typical technique for disposing of municipal solid waste [1]. Waste decomposition and residual water in landfills generate significant leachate [2]. Landfill leachate contains high levels of organics and other pollutants, posing a threat to the environment and human health (Table S1 in Supporting information). Biotechnology is a promising and cost-effective wastewater treatment technique [3]. However, such method is limited by the high concentration of bio-refractory pollutants (e.g., humic acid (HA), perfluorooctanesulfonate), heavy metals, ammonium (NH₄⁺), and salt [4–7]. Advanced oxidation processes (AOPs) such as Fenton, photocatalysis, and ozone catalysis can disrupt the molecular structure of refractory organics in landfill leachate [8–12]. Although AOPs can generate hydroxyl (^•OH) or sulfate radicals (SO₄^•-) to convert organic ammonium into NH₄⁺, NH₄⁺ removal remains a significant challenging [13]. Besides, Fenton or heterogeneous catalytic oxidation processes generate excess sludge, increasing treatment costs, necessitating the development of effective alternative methods.

  Based on landfill age, leachate can be classified into young, intermediate, and aged stages. The chemical oxygen demand (COD) concentration of young and intermediate leachate is typically above 4000 mg/L and can even exceed 10,000 mg/L, and the organics contain higher levels of low molecular weight organic acids and aliphatic organic compounds [14]. The ratio of 5-day biochemical oxygen demand to chemical oxygen demand (BOD₅/COD) is commonly above 0.4, indicating that biological method is more suitable [15]. As landfill age increases, low molecular weight organic acids are removed, leaving behind large, refractory compounds such as HA and antibiotics. Although COD drops below 4000 mg/L, the BOD₅/COD ratio remains below 0.15 [15], inhibiting the biological activity. However, recalcitrant organic compounds can be effectively degraded by electrochemical oxidation (EO), making this method a promising alternative for treating leachate from aged landfills.

  EO of landfill leachate has gained increasing attention for its effectiveness in degrading recalcitrant organics, removing nitrogen, and producing no secondary pollution. Organics can directly lose electrons on anode surface and simultaneously decomposes into smaller molecule substances [16]. NH₄⁺ can be directly oxidized on noble metal (such as Pt, Pd) anode surface [17]. H₂O, Cl^-, and SO₄^2- in leachate can be converted into reactive species such as ^•OH, Cl^•, and SO₄^•-, which degrade complex humic and fulvic acids into phenolic residues, aniline derivatives, and polycyclic aromatic hydrocarbons, eventually mineralizing them into CO₂ [18]. NH₄⁺ can be oxidized to N₂ by Cl^•, which is the main pathway for nitrogen removal in EO [19].

  The direct electron transfer process and indirect oxidation by oxidants primarily depend on the anode for the electrochemical reaction. Recently, various anodes such as boron-doped diamond (BDD), lead dioxide (PbO₂), ruthenium dioxide (RuO₂), tin dioxide (SnO₂) and their composite electrode were studied for EO landfill leachate treatment. For example, 83.7% of COD was removed under 100 mA/cm² with ruthenium iridium oxide electrode (Ti/RuO₂-IrO₂) [20]. The use of graphite/PbO₂ electrode could remove 35.1% COD and 97.6% NH₄⁺ from leachate after 2 h of electrolysis [21]. However, studies on electrochemical landfill leachate treatment are still largely limited to the laboratory stage due to technical limitations of the electrodes, including poor electron transfer performance, small specific surface area, single reactor operation mode, and insufficient space utilization.

  To overcome these limitations and enhance the electrooxidation efficiency of landfill leachate, extensive research efforts have been undertaken. However, current reports on EO mechanisms are mostly generic, lacking theoretical studies tailored to the characteristics of landfill leachate. Meanwhile, recent review papers on EO have mostly focused on comparing technologies like electrocoagulation, electro-adsorption, electro-Fenton, and photoelectrocatalysis [3,22,23], with scant attention given to enhancing the anodic electrooxidation efficiency. Therefore, this paper specifically summarized the EO mechanisms applicable to treating leachate from aged landfills, proposed feasible strategies to increase the electrode activity, and discussed the influence mechanism of operating parameters. Its goal is to provide guidance for the practical EO engineering of landfill leachate.

  2. Mechanisms

  The removal of pollutants in landfill leachate using EO technology can be categorized into direct and indirect pathways (Fig. 1). The direct oxidation refers to the capture of electrons from contaminants by active sites on anode under applied voltage [24]. Such reaction is a straightforward oxidation process. Indirect oxidation is the breakdown or mineralization of pollutants by oxidizing active chemicals (OACs) (^•OH, SO₄^•-, active chlorine, etc.) generated from anode [25].

  Figure 1

  

  Figure 1.  Schematic illustration for the EO process. R denotes an organic pollutant and M is the active site on the electrode surface.

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  2.1 Direct oxidation

  The direct oxidation process contains two steps. (a) Pollutants diffuse from the electrolyte to active sites on anode surface. (b) The electrode takes electrons from pollutants (Eq. 1) since the active sites have a high capacity for acquiring electrons [26]. The direct oxidation commonly has easy electron transport and quick reactions.

  $ \mathrm{M}+\mathrm{R} \rightarrow \mathrm{M}(^•\mathrm{R})+\mathrm{ne}^{-} \rightarrow \text { Product } $

  (1)

  M is the active site. R is the target pollutant.

  But, for aged leachate, the organic compounds often contain stable molecular structures such as benzene rings, naphthalene rings, heterocycles, which are difficult to degrade through direct oxidation pathways [16]. For example, the free radical quenching experiment confirms that direct oxidation only achieved a removal efficiency of 39.6% for HA [27]. More importantly, pollutants in leachate that cannot be quickly decomposed may accumulate on the electrode surface, leading to passivation. Therefore, degradation of leachate primarily relies on indirect oxidation pathways.

  2.2 Indirect oxidation mechanism

  2.2.1   The mechanism of OACs generation and transformation

  The indirect oxidation is mediated by various OACs including free radicals and reactive substances generated from anode. Generally, water can be decomposed into ^•OH (2.8 V vs. SHE) (Eq. 2) in electrochemical system, which has the strongest oxidizing characteristics [28]. The ^•OH also has high electronegativity or electrophilicity (the electron affinity is 569.3 kJ/mol), tending to attack sites with high electronic density in organics [29]. The ^•OH is produced and adsorbed on the anode surface, where it is employed for organic oxidation (Eq. 3) via electron transfer, dehydrogenation, and addition [30]. The organics can be converted into small molecule organics or CO₂ through such reaction. However, ^•OH can also cause oxygen evolution reaction (OER) with H₂O (acid condition) or OH^- (alkaline condition) and then release electrons (Eqs. 4 and 5) [31]. This side reaction will impair the efficiency of pollutant oxidation and increase energy consumption.

  $ \mathrm{M}+\mathrm{H}_2 \mathrm{O} \rightarrow \mathrm{M}\left({ }{^•\mathrm{OH}}\right)+\mathrm{H}^{+}+\mathrm{e}^{-} $

  (2)

  $ \mathrm{M}({^•\mathrm{OH}})+\mathrm{R} \rightarrow \mathrm{M}+\mathrm{RO}+\mathrm{H}^{+}+\mathrm{e}^{-} $

  (3)

  $ \mathrm{M}({^•\mathrm{OH}})+\mathrm{H}_2 \mathrm{O} \rightarrow \mathrm{M}+\mathrm{O}_2+3 \mathrm{H}^{+}+3 \mathrm{e}^{-} $

  (4)

  $ \mathrm{M}\left({^•\mathrm{OH}}\right)+\mathrm{OH}^{-} \rightarrow \mathrm{M}+\mathrm{O}_2+2 \mathrm{H}^{+}+3 \mathrm{e}^{-} $

  (5)

  In addition, Cl^-, a common anion in landfill leachate, can be oxidized into Cl^• through dimensional stability anode (DSA), such as RuO₂-Ir, IrO₂-TiO₂, and Co₃O₄ [32]. These DSA can enhance electron transfer and catalytic activity through the cyclic changes in the valence states of their constituent metal elements, enabling direct oxidation of Cl^- to Cl^• at the active sites by electron extraction near the anode. For example, the Co²⁺/Co³⁺ redox couple in the Co₃O₄ anode can act as a catalyst pair to rapidly oxidize Cl^- to Cl^•, enabling efficient removal of NH₄⁺ [33]. The Cl^- can also be indirectly oxidized to Cl^• by ^•OH (Eq. 6) [3]. The Cl^• will further react to form various active chlorine species (Eqs. 7-9), such as Cl₂ (E⁰ = 1.36 V vs. SHE), HClO (E⁰ = 1.63 V vs. SHE), and ClO^- (E⁰ = 0.90 V vs. SHE) [34,35]. Notably, active chlorines exist in different forms under different pH, such as Cl₂ (pH ≤ 3), HClO (pH 5–7), ClO^- (pH ≥ 8) [36]. However, electrooxidation mediated by active chlorine might produce some toxic chlorination byproducts like chloramine, trihalomethanes, haloacetic acids [30,37].

  $ {^•\mathrm{OH}}+\mathrm{Cl}^{-} \rightarrow \mathrm{OH}^{-}+\mathrm{Cl}^• $

  (6)

  $ 2 \mathrm{Cl}^{•} \rightarrow \mathrm{Cl}_2 $

  (7)

  $ \mathrm{Cl}_2+\mathrm{H}_2 \mathrm{O} \rightarrow \mathrm{HOCl}+\mathrm{H}^{+}+\mathrm{Cl}^{-} $

  (8)

  $ \mathrm{HOCl} \rightarrow \mathrm{ClO}^{-}+\mathrm{H}^{+} $

  (9)

  The generation and action mechanisms of SO₄^•- are shown in Fig. 2. The SO₄^2- and HSO₄^- in leachate can be directly oxidized to SO₄^•- on anode surface through single-electron transfer [38]. The potential for the direct conversion of SO₄^2- or HSO₄^- to SO₄^•- is typically around 2 V vs. Ag/AgCl [39]. Thus, such direct oxidation pathway is very weakly [40]. The SO₄^•- is mainly obtained from the indirect oxidation process mediate by ^•OH during electrochemical leachate treatment [41]. Meanwhile, the yield and stability of SO₄^•- are significantly dependent upon the pH. Low pH (pH ≤ 3), ^•OH can maintain excellent oxidation ability, promoting the generation of SO₄^•- (Eq. 10) [42]. The SO₄^•- has good stability and its hydrolysis can be ignored under such condition [43]. In contrary, SO₄^•- easy hydrolyzes into ^•OH at pH 9.0 (Eqs. 11 and 12). Remarkably, SO₄^•- will be completely react with OH^-/H₂O to ^•OH (Eqs. 11 and 12) as pH ≥ 12. The types of active species present in a system can be identified through radical probes, electron paramagnetic resonance (EPR), or quenching experiments (Text S1, Tables S2 and S3 in Supporting information).

  $ \mathrm{SO}_4^{2-}+{^•\mathrm{OH}} \rightarrow \mathrm{SO}_4 ^{•-}+\mathrm{OH}^{-} $

  (10)

  $ \mathrm{SO}_4{ }^{•-}+\mathrm{OH}^{-} \rightarrow \mathrm{SO}_4{ }^{2-}+{^•\mathrm{OH}} $

  (11)

  $ \mathrm{SO}_4{ }^{•-}+\mathrm{H}_2 \mathrm{O} \rightarrow \mathrm{HSO}_4^{-}+ {^•\mathrm{OH}} $

  (12)

  Figure 2

  

  Figure 2.  Sulfate radical-mediated oxidation and anode-cathode synergy.

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  The OACs can also produce from the synergistic performance of oxidation and reduction during electrochemical reaction. For example, the oxygen reduction reaction (ORR) can selectively reduce O₂, either generated at the anode or dissolved in water, to H₂O₂ (Eqs. 13 and 14). Specifically, O₂ is first adsorbed onto active sites and then, with the involvement of protons (H⁺), is ultimately reduced to H₂O₂ via a two-electron pathway on the cathode surface [44]. Since the reaction requires the protons, it occurs to a greater extent in young and intermediate-aged landfill leachate compared to aged leachate. H₂O₂ can also be formed through the dimerization of two ^•OH radicals (Eq. 15). Similarly, SO₄^2- is oxidized at the anode surface through a direct oxidation process or an indirect oxidation mediated by ^•OH, generating SO₄^•-, which can further dimerize into persulfate (S₂O₈^2-) (Eq. 16) [45]. The H₂O₂ and S₂O₈^2- can be reactivated at the cathode via a direct electron transfer pathway, producing ^•OH and SO₄^•- (Eqs. 17 and 18), respectively [45].

  $ 2 \mathrm{H}_2 \mathrm{O} \rightarrow \mathrm{O}_2+4 \mathrm{H}^{+}+4 \mathrm{e}^{-} $

  (13)

  $ \mathrm{O}_2+2 \mathrm{H}^{+}+2 \mathrm{e}^{-} \rightarrow \mathrm{H}_2 \mathrm{O}_2 $

  (14)

  $ 2 {^•\mathrm{OH}} \rightarrow \mathrm{H}_2 \mathrm{O}_2 $

  (15)

  $ 2 \mathrm{SO}_4{ }^{•-} \rightarrow \mathrm{S}_2 \mathrm{O}_8^{2-} $

  (16)

  $ \mathrm{H}_2 \mathrm{O}_2+\mathrm{e}^{-} \rightarrow{ }^{•} \mathrm{OH}+\mathrm{OH}^{-} $

  (17)

  $ \mathrm{S}_2 \mathrm{O}_8{ }^{2-}+\mathrm{e}^{-} \rightarrow \mathrm{SO}_4{ }^{•-}+\mathrm{SO}_4{ }^{2-} $

  (18)

  H₂O₂ or S₂O₈^2- can also be activated by indirect pathways (Fig. 2). The ORR reaction at the cathode can also produce reactive oxygen substances (O₂^•- and HO₂^•) (Eq. 19). The O₂^•-, as a highly unstable substance, can combine with H⁺ to form HO₂^• (Eqs. 20 and 21). Although O₂^•- and HO₂^• with low redox potential have poor oxidation property for pollutants removal, they can effectively oxidize H₂O₂ and S₂O₈^2- to ^•OH and SO₄^•- by seizing electrons (Eqs. 22-25), respectively [45,46].

  $ \mathrm{O}_2+\mathrm{e}^{-} \rightarrow \mathrm{O}_2^{•-} $

  (19)

  $ \mathrm{O}_2^{•-}+\mathrm{H}^{+} \rightarrow \mathrm{HO}_2^{•} $

  (20)

  $ \mathrm{OH}^{-}+\mathrm{HO}_2^{•} \rightarrow \mathrm{O}_2^{•-}+\mathrm{H}_2 \mathrm{O} $

  (21)

  $ \mathrm{H}_2 \mathrm{O}_2+\mathrm{O}_2^{•-} \rightarrow \mathrm{O}_2+\mathrm{OH}^{-}+{^•\mathrm{OH}} $

  (22)

  $ \mathrm{H}_2 \mathrm{O}_2+\mathrm{HO}_2^{•} \rightarrow \mathrm{O}_2+\mathrm{H}_2 \mathrm{O}+{^•\mathrm{OH}} $

  (23)

  $ \mathrm{S}_2 \mathrm{O}_8^{2-}+\mathrm{HO}_2^{•} \rightarrow \mathrm{O}_2+\mathrm{SO}_4^{2-}+\mathrm{SO}_4^{•-}+\mathrm{H}^{+} $

  (24)

  $ \mathrm{O}_2{ }^{•-}+\mathrm{S}_2 \mathrm{O}_8{ }^{2-} \rightarrow \mathrm{SO}_4{ }^{•-}+\mathrm{SO}_4{ }^{2-}+\mathrm{O}_2 $

  (25)

  2.2.2   Oxidation mechanism of leachate

  Taking the above discussion into account, ^•OH, SO₄^•- and Cl derivatives may be the primary active agents for leachate treatment. Due to the unique property leachate, its degradation pathway mainly involves the removal of organics and NH₄⁺. The SO₄^2- is a common ion in landfill leachate, which can be oxidized to SO₄^•- through direct or an ^•OH-mediated indirect oxidation process [41]. The SO₄^•- can degrade aromatic contaminants in landfill leachate. The UV–vis and three-dimensional fluorescence spectroscopy analysis showed a decrease of about 65% in aromaticity and hydrophobicity, and a 49.03% reduction in humification after SO₄^•- oxidation of landfill leachate in the Fenton process [47]. However, in EO process, SO₄^•- can be quenched by ^•OH. For example, SO₄^•- is not observed in the EPR spectra during the electrooxidation of HA [48,49]. Thus, organics and NH₄⁺in leachate are mainly removed by the ^•OH and Cl derivatives.

  Compared with ^•OH, Cl derivatives show significantly faster degradation rates of pollutants, especially for organics. For example, Cl derivatives exhibited 1.9 times higher efficiency than that of ^•OH in decomposing non-purgeable organic carbon and faster color removal from dark brown landfill leachate [48]. This is because ^•OH accumulates at active sites on anode surface through physical adsorption [30]. To undergo ^•OH mediated oxidation reaction, pollutants must transfer from solution to anode surface. Inversely, Cl derivatives generated on anode are soluble in solution and diffuse throughout the entire reactor, expanding the effective reaction area and enhancing removal efficiency. On the other hand, Cl derivatives can rapidly break down large molecules in leachate into smaller compounds such as phenol, aniline, benzoic acid, and polycyclic aromatic carbon hydride [50,51]. However, their lower oxidation potentials (Cl₂ 1.36 V vs. SHE, HClO 1.63 V vs. SHE, and ClO^- 0.90 V vs. SHE) not only make it difficult to further degrade the stable aromatic ring structures, but also result in the formation of chlorinated aromatic compounds [35]. The ^•OH with higher oxidation potential facilitates dechlorination and hydroxylation of these chlorinated products [52,53]. Research suggests that hydroxylation is a pivotal step in the aromatic ring-opening process, thus ^•OH achieves much higher mineralization efficiency for organic compounds in leachate [28,54]. However, the redox potential and stability of ^•OH is essentially dependent on the solution pH. For example, ^•OH can be quenched by H₂O or OH^- under neutral or alkaline conditions (Eqs. 4 and 5). Strong alkaline conditions can also decrease the redox potential of ^•OH, weakening the oxidative ability [55]. It is worthy nothing that HA is the most important contaminant in aged leachate, accounting for >60% of total organics. And its existing form is significantly affected by the solution pH. The detailed influence mechanism of pH will be further discussed in the section of 4.3.

  NH₄⁺, especially in aged leachate, can exceed 1500 mg/L [15]. Active chlorine can selectively oxidize NH₄⁺ to N_2, which is the primary pathway for NH₄⁺ removal in landfill leachate (Eqs. 26-31) [56]. For example, active chlorine achieved a removal efficiency of 97.85% for NH₄⁺ in landfill leachate after 8 h of EO, while ^•OH only removed 81.18% [57]. Taking the aged leachate (weakly alkaline) into account, active chlorine predominantly exists in the form of ClO^- (E⁰ = 0.90 V vs. SHE). ClO^- oxidation of NH₄⁺ can minimize the formation of by-products such as nitrite (NO₂^-) and nitrate (NO₃^-) [58].

  $ \mathrm{HOCl}+\mathrm{NH}_4^{+} \rightarrow \mathrm{NH}_2 \mathrm{Cl}+\mathrm{H}_2 \mathrm{O}+\mathrm{H}^{+} $

  (26)

  $ \mathrm{HOCl}+\mathrm{NH}_2 \mathrm{Cl} \rightarrow \mathrm{NHCl}_2+\mathrm{H}_2 \mathrm{O} $

  (27)

  $ \mathrm{H}_2 \mathrm{O}+\mathrm{NHCl}_2 \rightarrow \mathrm{NOH}+2 \mathrm{H}^{+}+2 \mathrm{Cl}^{-} $

  (28)

  $ \mathrm{NOH}+\mathrm{NHCl}_2 \rightarrow \mathrm{~N}_2+\mathrm{HOCl}+\mathrm{H}^{+}+\mathrm{Cl}^{-} $

  (29)

  $ 2 \mathrm{NH}_4^{+}+3 \mathrm{HOCl} \rightarrow \mathrm{~N}_2+3 \mathrm{H}_2 \mathrm{O}+5 \mathrm{H}^{+}+3 \mathrm{Cl}^{-} $

  (30)

  $ \mathrm{NH}_4^{+}+4 \mathrm{HOCl} \rightarrow \mathrm{NO}_2^{-}+\mathrm{H}_2 \mathrm{O}+6 \mathrm{H}^{+}+4 \mathrm{Cl}^{-} $

  (31)

  Overall, Cl derivatives significantly contribute to the breakdown of macromolecules (e.g., HA), NH₄⁺ removal, and decolorization, while ^•OH plays a key role in the complete mineralization of organic pollutants in aged leachate.

  3. Challenges and countermeasures

  Based on the above discussion, the primary pathway for removing organics and NH₄⁺ from landfill leachate is indirect oxidation mediated by OACs. The generation of OACs is closely linked to the anode. How to design the anode with excellent performance for steady and efficient production of active chemicals is the main focus of the recent of electrochemical landfill leachate treatment. Generally, the performance of anode is related on its physical and electrochemical characteristics including mechanical stability, conductivity, active sites, and mass transfer efficiency. To promote and regulate anode property, several approaches have been proposed (Table S4 in Supporting information), including adding intermediate layer materials, synthesizing porous membrane structures, creating surface microstructures, and filling particle microelectrodes.

  3.1 Introducing multifunctional intermediate layer

  The common anode was constructed with catalyst layer and a conductive base. However, catalyst layer (β-PbO₂, RuO₂, and IrO₂, etc.) has poor adherence to substrate, easy slipping off from electrode during electrochemical reaction, particularly under high current density. Moreover, the catalyst layer coverage on the substrate is often incomplete, leading to passivation of the base. Introducing an active intermediate layer can effectively address these challenges (Fig. 3).

  Figure 3

  

  Figure 3.  Diagram of the anode with intermediate layer.

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  The intermediate layer can enhance material adhesion and reduce substrate exposure [59]. Metal oxide or metal element not only have good adhesion, but also own high conductivity. For instance, Pt, with its excellent adhesion and conductivity, can serve as the intermediate layer to enhance electrode stability and reduce electrochemical impedance [60]. The electrode with SnO₂-Sb intermediate layer (Ti/SnO₂-Sb/β-PbO₂) features a long service life (214 h) and low electrochemical impedance (1.315 Ω/cm²) [61]. Besides, the intermediate layer can regulate the crystal of catalysts structure, improving the performance of electrode. The introduction of siloxides (SiO_x), a hydrophilic material, can promote the nucleation of β-PbO₂ crystals and further reduce electrochemical impedance [62].

  The intermediate layer material with special morphological characteristics such as two-dimension, porous and nanowire can improve the anode’s performance. The special morphological characteristic provides more sites for catalyst loading and bonding [63]. For instance, the introduction of MXene (Ti₃C₂T_x) with excellent conductivity, layered structure and hydrophilicity can significantly increase the specific surface area of the anode. The number of active sites and ^•OH output of Ti/Ti₃C₂T_x/PbO₂ were 5.21 and 4.07 times higher than Ti/PbO₂, respectively [64]. Besides, carbon nanotubes (CNTs) or other nanomaterial can be regarded as intermediate layer to increasing the specific surface area of anode and improving improve the adhesion between each layer, enhancing the electrochemical performance for landfill leachate treatment [65]. However, although carbon-based materials offer excellent electrochemical performance and morphological versatility, the high potential and acidic environment can lead to carbon oxidation and performance degradation over time [66].

  3.2 Designing porous membrane electrode

  The EO of leachate is achieved via indirect oxidation with free radical. However, free radical exists only in the boundary layer of micrometer or even nanometer level on electrode surface. The traditional operating mode limits the mass transfer efficiency between pollutants and the anode surface [67]. Flow-through membrane electrodes (MEs) with circular flakes and hollow cylindrical shapes have been developed (Fig. 4). These structures not only increase the electrochemically active sites but also enhance the diffusion of contaminants onto the electrode surface [48,68,69]. Researchers found that flow-through membrane electrode can reduce the thickness of diffusion layer from 100 µm to 1 µm, breaking the shackles of mass transfer by-pass operation [70]. Contaminants can be degraded and even mineralized within 5 s with electrochemical membrane electrode in flow-through mode [71,72]. Titanium dioxide (Ti₄O₇), SnO₂, PbO₂, and RuO₂, after pressing and calcination, can form good porosity and are commonly used materials for fabricating porous membrane anodes [73–76].

  Figure 4

  

  Figure 4.  Diagram of the porous membrane anode in flow-through mode.

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  Doping metal elements (bismuth (Bi), antimony (Sb), niobium (Nb), iridium (Ir), etc.) into porous membrane materials can increase free electron density, reducing electrochemical impedance of electrode. For example, compared with SnO₂ porous membrane electrode (182 Ω), Sb-doped SnO₂ owns lower electrochemical impedance (62 Ω) [77,78]. Similarly, PbO₂-Bi and RuO₂-Ir materials have low electrochemical impedance compared to PbO₂ and RuO₂, respectively [79,80]. The doping metal element can also create oxygen vacancies and increase active sites through the substitution reaction, enhancing the electrocatalytic activity of electrode. For example, doping cerium (Ce) (1–3 at%) into Ti₄O₇ can significantly facilitate interfacial charge transfer and increase ^•OH production (37%−129%) due to the creating of surficial oxygen vacancies [68]. Anuchai et al. increased the oxygen vacancy concentration in SnO₂ by 19.5%, generating more ^•O₂^- radicals and thereby enhancing the oxidation efficiency of methyl orange from 58% to 98% [81]. Combining carbon-based materials with porous membrane electrodes can also effectively enhance anode activity by forming chemical bond interfaces [82]. Compared with pristine Ti₄O₇, the introduction of graphene oxide nanoparticles (GONs) significantly decreases charge-transfer resistance (from 73.87 Ω to 8.42 Ω) and generated ^•OH at 2.5–2.8 times higher rate due to the formation of GONs-O-Ti chemical bonds [83]. Li et al. also found that the generation of Fe-O-GO—O-Ti bonds between graphene oxide (GO), iron oxide (Fe₂O₃), and Ti₄O₇ can boost the electroactivation of peroxymonosulfate for 1,4-dioxane removal due to the increase of electron transfer rate and reactive sites [84].

  It is worth noting that landfill leachate contains a large amount of suspended organic matter, such as HA, which can lead to membrane fouling and significant deactivation of active sites when treated directly with membrane electrodes. Therefore, even though porous membrane electrodes exhibit high electrooxidation efficiency, they are more suitable for treating low-concentration leachate, such as effluent from biological or coagulation pretreatment processes.

  3.3 Regulating microstructure of catalyst

  In practical engineering for EO of leachate, plate electrodes with limited specific surface area and active sites are typically used. Recently, three-dimensional (3D) microstructured anodes (Fig. 5), such as nanotubes, nanowires, and nanorods, have gained attention. These anodes feature large specific surface areas and multiple active sites [85,86]. For instance, NH₄⁺ removal efficiency with cobalt oxide (Co₃O₄) nanowire was 2.58 times than that of ordinary Co₃O₄ plate [33].

  Figure 5

  

  Figure 5.  Diagram of anode with three-dimensional microstructure.

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  The catalyst with 3D microstructure exhibits excellent activity for pollutant removal. However, a key challenge is the simple and cost-effective integration of nanostructured catalytic layers with substrates. The common strategy is using polymer as binder to load electrocatalyst onto substrate. But the polymer inevitably increases impedance and affects stability. To solve these problems, researches have been committed to in-situ grow 3D catalyst onto the electrode surface (binder-free electrode). Blue-titanium nanoarrays (BTNAs) can be grown in-situ on Ti substrates through two-step methods involving anodic oxidation and cathodic reduction. Their multiple active sites and good electron transfer ability can significantly increase generation of free radicals [87]. Pierpaoli et al. prepared carbon nanostructured electrode through in-situ grown boron doped vertically aligned graphene walls (BCNWs) onto BDD interfacial layer. It was found that in landfill leachate electrooxidation, COD removal efficiency was 1.66 times than that of BDD with the same geometric area [88].

  Nanoscale modification of 3D catalytic layers is also an effective method to improve anode electrochemical activity. The loading of metal element (such as Pt, Rh, Pd) in the form of single atom or nanocluster can boost the oxidation capacity of the surrounding active sites. For example, the modification with carbon nanotubes supported Pt single atom can minimize oxygen evolution side reactions and boost active chlorine creation [89]. Additionally, introducing more oxygen-containing functional groups (e.g., hydroxyl, carboxyl, and carbonyl) into nanostructured carbon-based materials can alter the electron cloud density of surrounding atoms, expand the potential window, and enhance electrode performance [90].

  3.4 Filling particle microelectrode

  The reactive sites and electron transfer rate are the essential parameters for EO of landfill leachate. Generally, they remain constant for a given electrode, limiting the efficiency of leachate degradation. Researchers discovered that adding catalysts, referred to as particle microelectrodes, can improve EO efficiency (Fig. 6). This method not only expands the reaction space but also enhances electron transfer [91]. Carbon-based materials are commonly used as particle microelectrodes to enhance electron transfer and increase reactive sites [92]. Using carbon aerogel as particle microelectrodes to construct a three-dimensional reactor demonstrated enhanced and sustained activity for phenol removal, achieving an efficiency of up to 98% [93].

  Figure 6

  

  Figure 6.  Diagram of the packed bed reactor with particle microelectrodes.

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  Under electric field, particle microelectrodes will polarize to form numerous micro-electrolytic cells, serving as active centers for radical generation [94]. In an electro-Fenton system with Fe particles, ^•OH generation was enhanced, leading to improved COD removal [95]. Chen et al. studied EO of dinitrotoluene (DNT) wastewater assisted by Sn-Sb-Ag-modified ceramic particulates (SCP). The ^•OH content in a three-dimensional electrochemical system filled with SCP particles was 1.85 times higher than that without SCP [96]. Additionally, the use of filler particles provides multiple convergent-divergent channels, preventing blockages and ensuring stable EO of leachate [97]. Therefore, inserting particle microelectrodes into an electrochemical reactor is an effective method to improve landfill leachate treatment efficiency by increasing the electrocatalytic active area, thereby enhancing the removal of COD, NH₄⁺, and chromaticity [98].

  Precisely, enhancing pollutant mass transfer, increasing the specific surface area, and exposing more active sites are critical to improving anodic oxidation performance. Additionally, improving electron transfer efficiency within the anode material to enhance its conductivity can reduce energy consumption and prevent anodic electrocorrosion caused by excessive applied voltage. Therefore, innovative research on anode design and reactor construction should be conducted based on the above key points.

  3.5 Engineering applications

  Besides electrodes, many EO equipment are designed to achieve excellent leachate treatment efficiency. Typically, the equipment was designed with multiple sets of alternating anodes and cathodes, significantly increasing the reaction efficiency and decreasing hydraulic retention time [99]. The electric energy consumption of flow-through mode using porous membrane electrode is 6 times less than that of the traditional flow-by mode due to better mass transfer and more active sites [100]. In addition, the packed-bed composed of many micro-particles can greatly expand the reactive active sites. For example, compared with non-packed system, the packed-bed equipment achieved 1.25 times higher phenol oxidation efficiency at 40 mA/cm² for 2.72 h [97].

  In practical engineering, EO is often combined with other technologies to fully leverage its efficient degradation advantages and reduce treatment cost. For example, coagulation pretreatment reduced the COD of leachate from 2550 mg/L to 938 mg/L, and EO further degrades it to 88.4 mg/L. The energy consumption of “coagulation + EO” process is 62.26% of EO [101]. The combination of electro-Fenton (pretreatment) and EO for leachate treatment can reduce energy consumption from 218.75 (COD, EO) to 120 kWh/kg [102]. Besides, EO can be combined with cost-effective biological technologies. For instance, EO can decompose the biologically toxic organics in leachate, increasing the BOD₅/COD ratio from 0.04 to 0.37 [103]. The cost of EO for leachate treatment is 55.8 €/m³, while it is only 18.7 €/m³ for “EO + Sequencing Batch Biofilter Granular Reactor (SBBGR)” [104]. Therefore, EO should be combined with other technologies to achieve excellent efficiency and low energy cost in practical application.

  4. Effect of operating parameters

  EO performance in leachate is influenced by operating parameters such as current density, electrolyte type, pH, and temperature (Table S5 in Supporting information). These factors affect the type, concentration, longevity, and oxidation ability of active species, as well as the state and degradation pathways of pollutants (e.g., HA) [21,56,105–113].

  4.1 Current density

  At a given current density, substances in the solution can be activated by the active sites on the anode surface, generating OACs that degrade pollutants in the leachate [114]. Leachate EO is an electron transfer process; the higher the current density, the more intense the redox reactions. Based on the above reaction mechanism discussion, the influence pathway of current density on leachate treatment is involved in the OACs generation. In leachate oxidation, when the current density was increased from 30 mA/cm² to 80 mA/cm², free chlorine production increased by 2.5 times, resulting in an increase in organic nitrogen removal from 17 mg N L^-1 h^-1 to 66 mg N L^-1 h^-1 [115]. Kim et al. found that the generation of ^•OH at 30 mA/cm² is twice that of 10 mA/cm² using cobalt-promoted lead dioxide anode [116].

  However, excessive current density can trigger halogenation, leading to the formation of chlorinated organic compounds with significant biological toxicity, such as 1,2-dichloroethane, haloacetonitriles, and haloethanes [46]. Moreover, higher current density leads to increased energy consumption due to the enhancement of side reactions, such as oxygen or chlorine evolution (Eqs. 32-34) [117]. Using BDD anode to oxidize leachate, it was found that when the current density increased from 1 mA/cm² to 30 mA/cm², the effective utilization rate of electrical energy decreased from 100% to 57.8% [115]. Choosing an adequate current density is critical for encouraging leachate degradation, enhancing economic advantages, and avoiding secondary contamination. In conclusion, the selected range of current density is relatively broad (10–100 mA/cm²) and should be chosen based on the desired effluent standards and control of toxic by-products.

  $ \mathrm{HClO}+\mathrm{Cl}^{-}+\mathrm{H}^{+} \rightarrow \mathrm{Cl}_2+\mathrm{H}_2 \mathrm{O} $

  (32)

  $ 2 \mathrm{Cl}^{-} \rightarrow \mathrm{Cl}_2+2 \mathrm{e}^{-} $

  (33)

  $ 2 \mathrm{H}_2 \mathrm{O} \rightarrow 4 \mathrm{H}^{+}+\mathrm{O}_2+4 \mathrm{e}^{-} $

  (34)

  4.2 The component of electrolyte

  In EO, the electrolyte not only serves as the medium for electron transfer, but also provides the essential materials for the formation of OACs. Different electrolytes influence the types of active species and the degradation pathways of pollutants.

  The Cl^- in leachate is the key content for nitrogen removal. The Cl derivatives generated from Cl^- electrooxidation can quickly capture electrons from NH₄⁺ and simultaneously selectively convert it into N₂, achieving total nitrogen removal. Cl^• and its derivatives can remove 98.62% of NH₄⁺ at a concentration of 590 mg/L after 3 h at 125 mA/cm² [118]. Cabeza et al. found that when Cl^-:NH₄⁺ molar ratio exceeds 9:1, NH₄⁺ removal efficiency can reach over 88.51%. Below this ratio, NH₄⁺ removal efficiency significantly decreases [13]. Cifcioglu-Gozuacik et al. reported that when the Cl^-:NH₄⁺ ratio is 6:1, NH₄⁺ removal efficiency is only 40% [119]. The NH₄⁺ removal efficiencies are 96.5%, 78.8%, and 61.0% when the Cl^-:NH₄⁺ ratios are 9:1, 6:1, and 4.5:1, respectively [120]. The Cl^-:NH₄⁺ ratio above 9:1 is recommended to maintain a removal efficiency over 80%. Cl^- also plays an important role in leachate decolorization. This is attributed to the rapid reaction between Cl derivatives and the π-π bonds in chromophoric groups of polymeric organic compounds [121].

  The SO₄^2- is another anion in leachate. Theoretically, the SO₄^2- can be oxidized to SO₄^•- (Eqs. 16 and 35) through electron transfer from anode and ^•OH, promoting organic removal. For example, compared to NO₃^- electrolyte, the removal rate of dissolved organic carbon (DOC) in SO₄^2- increased from 41% to 51% through electrooxidation due to the generation of SO₄^•- [122]. However, SO₄^2- is an oxygen-containing ion that will induce oxygen evolution side reactions [113]. The hydrated ionic radius of SO₄^2- (2.40 Å) is larger than that of Cl^- (1.81 Å), enabling SO₄^2- to adsorb more easily onto the anode surface, occupying electroactive sites and inhibiting the formation of other active species [123,124].

  $ \mathrm{HSO}_4^{-}+{ }^{•} \mathrm{OH} \rightarrow \mathrm{SO}_4^{•-}+\mathrm{H}_2 \mathrm{O} $

  (35)

  Aged leachate also contains Ca²⁺ (200 mg/L) and Mg²⁺ (100 mg/L), which can easy form hydroxide precipitate under alkaline condition (pH ≥ 10.3) [125]. This precipitate will adhere to the anode surface, passivating the active sites. For example, 2,4-dichlorophenol removal efficiency decreased from 67.8% to 33.1% due to the passivation of active sites by Ca²⁺, Mg²⁺ [126]. The interface pH at cathode is as higher as 12 even when the solution pH is 7.7–7.9 [127]. To solve the influence of Ca²⁺ and Mg²⁺ on electrochemical process, several technologies have been developed. The crystallization fluidized bed effectively reduced the Ca²⁺ concentration from 537.81 mg/L to 78.43 mg/L [128]. The chemical precipitation controlled Mg²⁺ concentration below 10 mg/L [129]. Plate and frame filter press and electrodeposition are also efficient for removing Ca²⁺ and Mg²⁺.

  4.3 pH

  Leachate pH varies at different landfill stages. New leachate (0–5 years), rich in volatile fatty acids (VFAs) from anaerobic fermentation, has a pH < 6.5. As VFAs escape or biodegrade, the pH of intermediate leachate (5–10 years) rises to 6.5–7.5. Aged leachate (>10 years), with a pH > 7.5, primarily contains refractory organics like HA [14]. Leachate treatment with EO primarily relies on indirect oxidation mediated by OACs, whose generation is greatly influenced by pH.

  The generation of ^•OH is related to the pH. The oxygen evolution reaction (OER) is the main side reaction for ^•OH production, with its trigger potential decreasing as pH increases (Eqs. 36 and 37) [130]. The cyclic voltammetry (CV) demonstrates that OER redox peak shifts to lower potential at higher pH [131]. The high rate of oxygen evolution on electrode surface will occupy reactive sites for ^•OH generation, reducing the organics removal [132]. The high oxygen evolution potential (1.229 V vs. SHE) in acidic or neutral conditions will facilitate ^•OH generation [130,133], facilitating organics oxidation. In an EO experiment, the removal rate of chloramphenicol at pH 3 is about 2.4 times higher than at pH 11 [134]. Additionally, under acidic conditions, SO₄^•- is not quenched by hydrolysis (Eq. 31). At pH ≥ 9, SO₄^•- can react with OH^- to form ^•OH (Eq. 30), which then reacts with OH^- (Eq. 4) to produce oxygen, quenching OACs.

  $ 4 \mathrm{OH}^{-} \rightarrow 2 \mathrm{H}_2 \mathrm{O}+\mathrm{O}_2+4 \mathrm{e}^{-} $

  (36)

  $ 2 \mathrm{H}_2 \mathrm{O} \rightarrow 4 \mathrm{H}^{+}+\mathrm{O}_2+4 \mathrm{e}^{-} $

  (37)

  Active chlorine is key OACs for NH₄⁺ removal in leachate. At different pH levels, active chlorine exists as Cl₂ (pH ≤ 3), HClO (pH 5–7), ClO^- (pH ≥ 8) [36]. The NH₄⁺ removal rate is much higher in alkaline than in acidic solutions. Vlyssides et al. reported nearly 100% NH₄⁺ removal through EO at pH 9, compared to only 30% at pH 6 [135]. Importantly, although the active chlorine can directly oxidize NH₄⁺, the products depend on the reaction solution pH [135,136]. In acid condition, NH₄⁺ is easily oxidized to NO₃^- (by-product) by Cl₂, while it is selectively converted to N₂ by ClO^- under alkaline environment. For example, after 90 min of EO at pH 6, the NO₃^- concentration reaches 118.2 mg N/L from an initial NH₄⁺ concentration of 1094 mg N/L. In contrast, at pH 9, the NO₃^- concentration accumulates to only 7.2 mg N/L under the same conditions [135]. The pH of aged leachate typically ranges from 8 to 10, creating an ideal environment for NH₄⁺ removal through EO.

  The presence status of weak acid ion (CO₃^2- and HCO₃^-) and pollutants in leachate are also affected by the solution pH. The CO₃^2- and HCO₃^- are quenching agent for ^•OH, Cl^• and Cl₂^•-. They will be transferred into CO₂ under acidic or neutral conditions (pH < 8.3) especially pH < 4, improving the effective utilization of free radicals OACs. Besides, HA is the main organic pollutant in leachate particularly for aged leachate. It exhibits a flocculent state under acidic conditions (pH < 3.5), covering active sites and hindering pollutant diffusion [137]. For example, for real leachate with an initial COD of 42,000 mg/L, flocculation occurred at pH 6 and the corresponding COD removal efficiency reached 81.9% at pH 4.5 [138]. Besides, H⁺ induces protonation of carboxyl and phenolic groups in organic, leading to zeta potential increase and hydrophilicity decrease [139]. For example, at pH 2, zeta potential approached 0 mV and 86% of COD in leachate mainly removed by flocculation process [138]. At pH > 6.5, carboxyl and phenolic groups are in an ionic state and form opposite charge electric layers [140]. The eta potential is less than −14.7 mV, improving the flocs resolution and HA dispersion [138]. Therefore, EO for leachate degradation under neutral or weak alkaline conditions is a suitable option.

  4.4 Temperature

  High temperatures are conducive to electron transfer, mass transport, ionic strength and conductivity, improving the leachate treatment [56,141]. For example, raising the temperature from 35 ℃ to 55 ℃ improves the removal efficiency of total organic carbon (TOC) and NH₄⁺ from 59.7% and 88.8% to 70.2% and 99.9%, respectively [142]. Adjusting the temperature from 30 ℃ to 60 ℃, the decolorization rate increases from 40% to 70% [113]. Electrode activity is also significantly affected by temperature involving the electroactive area and current density [143]. When the temperature increased from 25 ℃ to 60 ℃, the current density of Pt electrode increased by 42.85% [143]. However, high temperature could also lead to an increase of by-products in electrochemical leachate treatment, especially in the presence of Cl^-. For example, the concentration of disinfection by-products (DBP) increased by 50% when the reaction temperature rose from 28 ℃ to 35 ℃ [144]. Perchlorate generation from Cl^- oxidation at 20 ℃ was 1.5 times higher than at 6 ℃ [145].

  Electric energy can also be converted into thermal energy, leading to a temperature increase during the EO process of leachate. Sun et al. found that the leachate temperature rose from 23 ℃ to 50 ℃ after electrolysis 3 h [142]. However, in practice, heating large volumes of leachate will increase costs and by-product formation. Therefore, it is recommended to conduct leachate EO at ambient temperatures (15–35 ℃) and utilize the inevitable temperature increase from electrical energy losses to heat the wastewater.

  5. Conclusions and prospects

  This review analyzed the main pathways for landfill leachate EO treatment and focused on the mechanisms of OACs generation and transformation for pollutant removal. In addition, recently developed countermeasures for enhancing electrode performance in landfill leachate treatment have been discussed. The introduction of a multifunctional intermediate layer between the substrate and catalyst improves the stability and conductivity of the anode. To accelerate mass and electron transfer, two approaches have been developed: designing porous membrane electrodes and incorporating particle microelectrodes. Regulating the catalyst microstructure can provide sufficient active sites for OACs generation and pollutant removal. Key parameters in the EO of leachate were also discussed to further optimize the application of this technology.

  However, technical barriers still limit the commercialization of each strategy. The following key areas for future research are proposed: (1) Research on EO for leachate degradation should shift from single pollutants (e.g., HA) to complex pollutant systems and from simulated wastewater to actual leachate, thereby advancing the development of high-performance anodes. The cost of electrode preparation and their actual service life need to be further evaluated. Additionally, the impact of Ca²⁺ and Mg²⁺ in leachate on electrode passivation, pollution control efficiency, and solution method needs further study. (2) Although EO can efficiently treat landfill leachate, it should be combined with other technologies in practice to increase treatment efficiency and reduce treatment costs. Additionally, the toxic by-products of leachate EO, especially their formation mechanisms and quantitative analysis, is limited. (3) Leachate contains various substance, leading to multiple active species during electrooxidation. Most researches focus on the removal of macroscopic indicators like COD/TOC, color, and NH₄⁺, with limited study on the degradation characteristics and pathways of specific pollutants. 4) Most studies on leachate EO have been conducted at the laboratory stage. More focus is needed on practical applications, particularly the design and development of EO equipment.

  Declaration of competing interest

  The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

  CRediT authorship contribution statement

  Zhenhui Song: Writing – review & editing, Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Data curation, Conceptualization. Xing Wu: Validation, Methodology, Investigation, Data curation. Tianyu Gao: Visualization, Methodology, Investigation. Fubing Yao: Writing – review & editing, Visualization, Validation, Investigation, Funding acquisition, Formal analysis, Conceptualization. Xi Tang: Validation, Methodology, Investigation, Conceptualization. Qaisar Mahmood: Writing – review & editing, Visualization, Validation. Chong-Jian Tang: Writing – review & editing, Validation, Supervision, Resources, Project administration, Funding acquisition, Formal analysis, Conceptualization.

  Acknowledgments

  This research was financially supported by the Innovative Research Groups of the National Natural Science Foundation of China (No. 52121004), the project of National Natural Science Foundation (No. U21A20294), the Scientific Foundation of Hunan Province (Nos. 2024RC1012, 2022JJ40622), The Technology Innovation Guidance Project of Jiangxi Province, China (Nos. 20203BDH80W017, 20212BDH81030).

  Supplementary materials

  Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.cclet.2025.111008.

  
  
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  Figure 1  Schematic illustration for the EO process. R denotes an organic pollutant and M is the active site on the electrode surface.

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  Figure 2  Sulfate radical-mediated oxidation and anode-cathode synergy.

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  Figure 3  Diagram of the anode with intermediate layer.

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  Figure 4  Diagram of the porous membrane anode in flow-through mode.

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  Figure 5  Diagram of anode with three-dimensional microstructure.

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  Figure 6  Diagram of the packed bed reactor with particle microelectrodes.

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