English

• Wastewater or seawater reclamation is an effective way to solve the water shortage problem, but the reverse osmosis membranes for pollutant and/or salt separation suffer inevitable biofouling [1]. Methylisothiazolinone (MIT) is a widely utilized non-oxidizing fungicide, renowned for its ability to provide enduring control over microbial risks without reacting adversely with other product components or causing equipment corrosion [2,3]. However, due to its inherent stability, MIT tends to accumulate in the reverse osmosis concentrate (ROC) water [1]. When released into the environment, MIT poses a significant threat to marine ecosystems because of its toxic nature and resistance to biodegradation [4,5]. Consequently, the issue of MIT pollution has garnered considerable attention and concern. Due to its biorefractory nature, traditional wastewater treatment processes are not efficient for the degradation [6]. Thus, it is of an urgent demand to develop effective technologies for elimination of MIT pollution in wastewater.

  Electrochemical advanced oxidation processes (EAOPs) have demonstrated exceptional efficiency, operational simplicity, and environmental friendliness in the transformation and degradation of persistent organic compounds within aqueous solutions [7]. These processes are increasingly recognized for their capacity to produce reactive oxygen species (ROS), which can effectively convert organic pollutants into less harmful, smaller molecules or even into completely mineralized forms [8]. In particular, three-dimensional (3D) electrode system has been attracting wide interest as an emerging EAOPs in recent years [9]. Notably, the 3D electrode system addresses several challenges faced by traditional two-dimensional (2D) electrochemical systems, including low space-time yield, mass transfer limitations, low current efficiency, low area-to-volume ratio, and temperature increases [10]. In a 3D electrochemical system, particle electrodes function as a third electrode, creating multiple micro-electrochemical cell reactors under the influence of an electric field [10]. In a 3D electrolysis process, both the anode and the particle electrodes can generate strong hydroxyl radical oxidant (^•OH) for chemical reactions, which can enhance the electrochemical reaction rate and improve the degradation rate of organic pollutants [11].

  Consequently, the adoption of 3D electrode technology has garnered increasing attention in recent years, highlighting its potential in the field of electrochemical treatment processes [12].

  Proper selection of the electrode material is a key strategy to ensure the effectiveness of the electrochemical oxidation process since the generation of oxidizing agents, such as hydroxyl radicals, hypochlorite, and ozone, is strongly dependent on the properties of the selected electrode [13]. Therefore, numerous efforts have focused on the development and design of high-efficiency electrode materials that exhibit high over potential for oxygen evolution reaction (OER) and high activity for producing oxidizing agents [13]. Particle electrodes are commonly considered to be crucial to the performance of electrocatalytic oxidation in 3D electrochemical system. In addition, it is found that the specific surface area of carbon-based materials is significantly higher than that of metal-based materials by comparing various 3D electrode materials [7]. Activated carbon is the most used particle electrodes in current research [14,15]. Activated carbon has the advantages of high mechanical strength, large specific surface area, excellent adsorption performance and long service life [16,17]. Granular active carbon (GAC) is easy to short-circuit because of its poor electron transfer capability and electrical reactivity [7] and its electrochemical oxidation capacity is poor as particle electrode in 3D system [10]. Therefore, it remains a challenge to develop the high-performance, low-cost particle electrodes for practical application of the 3D electrochemical system in wastewater treatment [7].

  Electrocatalytic oxidation treatment processes often rely on many key catalytic processes [18]. Several studies have investigated that metal-based catalysts often have better catalytic ability than other catalysts [19]. To augment catalytic efficacy, researchers coated a metal (Fe, Cu, Zn, Ti, Sn, Ce, Ni) oxide as a catalyst on the particle electrode [10]. Furthermore, the rare-earth element Ce as a dopant optimizes the electrocatalytic performance due to a high (Ce³⁺/Ce⁴⁺) redox potential and a synergistic effect between Ce and Sn, 2Ce⁴⁺ + Sn²⁺ → 2Ce²⁺ + Sn⁴⁺, which can enhance the redox capability and improve the electrocatalytic ability of the electrocatalyst [20]. However, the mechanism by which the 3D electrocatalytic system degrades MIT is not yet clear.

  Due to its high resistivity, strong adsorption, developed pores, and weak conductivity, this paper uses GAC as a particle electrode carrier for the 3D electrodes system to investigate using a sol-gel method to prepare Sn-Sb-Ce/GAC particle electrode to degrade MIT. Firstly, the optimum preparation conditions of Sn-Sb-Ce/GAC particle electrode were investigated by orthogonal experiment. The catalytic activity, oxidation capacity and stability of the modified particle electrode were investigated through the characterization of particle electrode, electrocatalytic oxidation and degradation of MIT, adsorption performance test and stability experiment. The effects of current density, initial pH, electrolyte concentration and reaction time on the removal rate of MIT were investigated by single factor method, and the active substances degrading MIT were analyzed. Finally, the 3D electrode and 2D electrode were compared to degrade MIT and pollutants in real ROC water.

  The 3D electrochemical operating devices in Supporting information and Fig. S1 (Supporting information). The electrocatalytic degradation efficiency of MIT by coconut shell activated carbon particle electrode was significantly higher than that of nut shell activated carbon particle electrode or coaly activated carbon particle electrode (Fig. S2a in Supporting information). Therefore, the coconut shell activated carbon was selected as the carrier of modified particle electrode. After electrocatalytic degradation for 2 h, the MIT removal rate of the 3D electrode system was superior to that of the 2D system, with an increase of at least 20% (Fig. S2b in Supporting information). Additionally, within the 3D electrode system, the electrocatalytic activity of the Sn-Sb-Ce/GAC particle electrode is superior to that of the GAC, Sn-Sb/GAC, and Sn-Sb-Ce/GAC particle electrodes (Fig. S2b). Therefore, based on the above results, this experiment selected coconut shell activated carbon as the carrier, Sn, Sb and Ce as the active substances, prepared composite supported particle electrode by sol-gel method, and formed a 3D electrode system for subsequent MIT degradation research.

  Orthogonal experiments were designed to optimize the parameters of particle electrode preparation. The results showed that the optimum conditions were obtained when the concentration of SnCl₄·5H₂O was 0.3 mol/L, the molar ratio of Sn and Ce was 50:3, the calcination temperature was 400 ℃, and the calcination time was 3 h (Supporting information).

  The scanning electron microscopy (SEM) images of GAC particle electrode loaded with three kinds of metallic elements were shown in Fig. 1. However, the SEM images of activated carbon loaded with Sn, Sb, Ce metals were significantly different. Fig. 1a showed that the blank GAC had a rough surface with many voids [16]. It could be clearly seen from Fig. 1b that the co-doping of the three components had a significant effect on the morphology of the electrode. The crystal particles were distributed uniformly on the surface of the electrode, which notably enhanced the roughness of the electrode, thereby potentially augmenting its catalytic activity and surface area for electrochemical reactions. Fig. 1c showed that compared with the bare activated carbon, the surface of Sn-Sb-Ce/GAC material had many spherical, cylindrical and cubic crystals of different shapes. Fig. 1c also indicated that the GAC surface and inner wall were unevenly covered with the Sn, Sb, Ce catalysts after the sol-gel treatment [10]. This phenomenon successfully demonstrates that the three kinds of metals were loaded onto the GAC particle electrode.

  Figure 1

  

  Figure 1.  (a) SEM of GAC (10 K). (b) SEM of Sn-Sb-Ce/GAC (10 K). (c) SEM of Sn-Sb-Ce/GAC (20 K). (d) EDS of Sn-Sb-Ce/GAC particle electrode. (e) XRD pattern of particle electrode.

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  The atom percentages of the Sn and Sb elements were approximately 10.54% and 0.23%, respectively, and the Ce element was approximately 0.75%, as obtained from Fig. 1d. The value was higher than the molar ratio of Sn: Sb = 10:1. This discrepancy suggested that the addition of Sb may have dispersed into the lattice of SnO₂, possibly altering its crystal structure, which in turn resulted in an increase in the atoms ratio of Sn: Sb [10]. Thus, it could be inferred that Sn-Sb-Ce/GAC materials were successfully synthesized.

  To delve into the crystal structure and crystallographic orientation of the GAC catalysts with Ce, Sn, and Sb doping, a comparative analysis of the X-ray diffraction (XRD) patterns were conducted [13]. As illustrated in Fig. 1e, the XRD patterns of GAC before and after loading were significantly different. Compared with the XRD spectrum of GAC, the XRD pattern of Ce/GAC was similar. The weak characteristic diffraction peaks at 2θ = 28.5° was attributed to the (111) reflection of cubic CeO₂ [21]. The energy-dispersive spectroscopy (EDS) proved the existence of the Ce element. However, unlike the Sn-Sb/GAC composite, the characteristic peak associated with CeO₂ was notably absent in the Ce-Sn-Sb/GAC system. This indicated that the Ce element was highly dispersed in the SnO₂ tetragonal crystal structure, and it manifested that Ce and Sn had a strong interaction, which was conducive to the increase of particle electrode conductivity and the improvement of electrocatalytic activity [20]. According to Fig. 1e, compared with Sn-Sb/GAC, the XRD of Ce-Sn-Sb/GAC showed obvious characteristic peaks at 2θ = 26.51°, 33.75° and 51.9°, which were attributed to the (110), (101) and (211) reflections of tetragonal SnO₂, respectively [22]. It belongs to the typical rutile structure oxide. The EDS demonstrated the existence of the Sn element. The intensity of the diffraction peak of SnO₂ was strong, indicating that SnO₂ had a high crystallinity. It can be seen from Fig. 1e that the peak position of the diffraction peak of SnO₂ after Ce was doped slightly shifted, and the intensity of the diffraction peak was weakened, indicating that Ce doping affected the growth of SnO₂ grains, and fewer reflective grains could be formed in the direction of the crystal axis of the same crystal [10]. Additionally, the absence of characteristic diffraction peaks for Sb₂O₅ at 2θ = 28.91° can be attributed to the formation of an amorphous phase of SnO₂-Sb at the relatively low calcination temperature of 500 ℃ [23]. In short, Sn, Sb and Ce were successfully loaded onto the GAC.

  The effects of current density, initial pH, electrolyte concentration and reaction time on the degradation of MIT in a 3D electrocatalytic system were studied in Figs. S3-S6 (Supporting information).

  Although mass transfer is one of the most crucial factors in the electrochemical degradation process, the physical and chemical adsorption of species on the electrode surface should not be ignored, as it is can significantly affect the electron transfer kinetics [7]. Therefore, the adsorption performance of the particle electrode is as important as its catalytic oxidation capacity. On the one hand, the particle electrode accumulates pollutants through adsorption and provides an active site for their degradation. The pollutants adsorbed on the surface of the particle electrode are more susceptible to oxidation and degradation by free radicals [24]. On the other hand, the particle electrode continuously adsorbs pollutants, increasing their concentration on the electrode surface above that in the liquid phase, thus accelerating the reaction. Therefore, the electrocatalytic oxidation performance of the particle electrode is greatly affected by its adsorption performance, with strong adsorption abilities being conducive to the degradation of pollutants. The adsorption test was carried out with reference in Supporting information. The adsorption efficiency of the Sn-Sb-Ce/GAC particle electrode decreased with the increase of the dose (Fig. S7 in Supporting information). When the particle electrode was 5 g/L, the adsorption efficiency was at its maximum (Fig. S7a). Hence, the 3D electrode system optimal reaction condition: [MIT] = 50 mg/L, current density = 20 mA/cm², initial pH 5, [Na₂SO₄] = 0.2mol/L, 120 min, particle electrode was 5 g/L.

  The stability and reusability of the particle electrode are crucial factors in determining its engineering application [24]. The removal efficiency of MIT decreased from 92.93% to 88.49% after ten cycles, but there was no significant decrease in the MIT removal rate (Fig. S7b). The low dissolution amount of Ce, Sn and Sb ions ensured the long-term effectiveness of the catalytic oxidation capacity of the particle electrode and would not cause secondary pollution of the water body (Table S3 in Supporting information). In conclusion, the structural stability of the Sn-Sb-Ce/GAC particle electrode has a certain persistence and has the potential of engineering applications.

  The degradation of organic pollutants by 3D electrode system can be divided into direct oxidation and indirect oxidation [25]. Direct oxidation is characterized by easy electron transfer and rapid reaction, but it has relatively low oxidizing power. Indirect oxidation is mediated by various oxidizing active chemicals (OACs), including free radicals and reactive substances generated from the anodes.

  To further elucidate the degradation mechanism of MIT in Sn-Sb-Ce/GAC particle electrode system, experiments were conducted using tertiary butanol (TBA) and methyl alcohol (MeOH) as probe molecules for different ROS. The contribution ratio of the corresponding ROS can be evaluated based on the inhibitory degree caused by the introduction of a scavenger [26].

  TBA is commonly used as a quencher for testing the contribution of ^•OH to the degradation of contaminants due to its high sensitivity to ^•OH. While MeOH can quench both ^•OH and ^•SO₄^-, an appropriate amount of MeOH is added to the system to quench these radicals. As shown in Fig. 2a, at the end of the reaction, the degradation rate of MIT with TBA was 73.9%, which was lower than degradation rate of 93.9% when TBA was not be added. The 20.0% decrease in removal rate indicated that the contribution of ^•OH to the oxidation degradation of MIT was 20%. Meanwhile, the reaction rate constant without TBA was 1.5 times higher than that with TBA. The inhibitory effect of TBA on MIT degradation indicated that indirect oxidation by reactive oxygen species played a significant role in the electrocatalytic reaction of Sn-Sb-Ce/GAC 3D electrode system.

  Figure 2

  

  Figure 2.  (a) The effect of TBA and MeOH on the removal effect of MIT and the curve of degradation kinetics. (b) The amount of active chlorine produced during electrocatalytic oxidation. (c) Contribution rate of different oxidation modes in the electrocatalytic oxidation of MIT by 3D electrode system. The 3D electrode system reaction condition [MIT] = 50 mg/L, current density = 20 mA/cm², initial pH 5, [Na₂SO₄] = 0.2 mol/L, Sn-Sb-Ce/GAC particle electrode was 5 g/L.

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  As can be seen from Fig. 2a, the treatment efficiency of the system was greatly reduced after the MeOH was added. After 120 min of electrocatalytic degradation, the removal rate of MIT was only 17.2%, indicating that the direct oxidation contribution of the system was 17.2%. At the same time, the reaction rate constant of the system without MeOH was 14 times that of the system with MeOH. In addition, the reaction rate of direct oxidation increases gradually, indicating that the degradation of pollutants by the anode is relatively stable, and the treatment efficiency does not fluctuate greatly.

  Indirect oxidation involves the decomposition or mineralization of pollutants by OACs generated at the anode, such as hydroxyl radicals (^•OH), sulfate radicals (SO₄^•-), and active chlorine [27]. In the 3D electrode system, since the operating potential is lower than that required for water electrolysis, the destruction of pollutants occurs through direct oxidation. Conversely, if the system potential exceeds this value, OACs will form on the anode surface, leading to indirect oxidation.

  Under normal circumstances, water can be decomposed into ^•OH (2.8 V vs. SHE) (Eq. 1) in an electrochemical system, which exhibit the strongest oxidizing characteristics [28]. Here, M represents the active site, and R denotes the target pollutant.

  M + H2O → M(•OH) + H+ + e-

  (1)

  Besides, ^•OH has a high electronegativity and tends to attack sites with high electronic density in organics. The ^•OH is generated and adsorbed on the anode surface, where it undergoes organic oxidation (Eq. 2) through electron transfer, dehydrogenation, and addition reactions [24].

  M(•OH) + R → M + RO + H+ + e-

  (2)

  Then, organic compounds are transformed into small molecule organics or CO₂ through free radical chain reactions [29,30]. However, ^•OH can also cause oxygen an evolution reaction (OER) with H₂O under acidic conditions, releasing electrons (Eq. 3) [25]. This side reaction can impair the oxidation efficiency of pollutants and increase energy consumption.

  M(•OH) + H2O → M + O2 + 3H+ + 3e-

  (3)

  Because the Sn, Sb, Ce coatings compensate for the deficiencies of GAC, promoting the electrocatalytic oxidation reaction on the surface of the particle electrode [31,32]. Therefore, the degradation rate of MIT is increased. In addition, for the porous structure of Sn-Sb-Ce/GAC, the multiple reaction sites and hydrophobicity reduce the H₂O adsorption at the anode surface, hindering the expansion and movement of ^•OH towards the interior of the electrode, thereby suppressing side reactions such as OER and enhancing the oxygen evolution potential (OEP) of electrode [33,34].

  On the other hand, the contribution of active chlorine to MIT degradation was studied while maintaining the charge equality of the electrolyte. ROC wastewater usually contains a high concentration of chloride ions. The high concentration of chloride ions can be converted into strong oxidizing agents such as Cl₂, HOCl and OCl^- in the electrocatalytic oxidation system, thus promoting the oxidative degradation of pollutants. Anodic oxidation of chloride ions to chlorine is the primary pathway for generating active chlorine (Eqs. 4–6) [24]. The Cl^- will be derived into variety forms of active chlorine, such as Cl₂ (E⁰ = 1.36 V vs. SHE), HClO (E⁰ = 1.63 V vs. SHE), and ClO^- (E⁰ = 0.90 V vs. SHE) [24,28]. Notably, active chlorine exits in different forms under different pH, such as Cl₂ (pH ≤ 3), HClO (pH 5–7), ClO^- (pH ≥ 8) [35].

  2Cl- → Cl2 + 2e-

  (4)

  2Cl2 + H2O → HClO + H+ + Cl-

  (5)

  HClO ↔ H+ + ClO-

  (6)

  According to Fig. 2b, it can be seen that after 120 min of reaction, the concentration of active chlorine in the 3D electrocatalytic oxidation process reached 1640.37 mg/L. As the reaction proceeded, the generation rate of active chlorine also gradually slowed down. Fig. 2c indicated that there are three oxidation mechanisms in the 3D electrode system: direct oxidation, active chlorine oxidation, and hydroxyl radical oxidation. The contribution of reactive species generated by indirect oxidation to MIT degradation were significantly higher than those by direct oxidation. Active chlorine played a dominant role in the degradation of MIT.

  The pH of the 3D electrode system is 5, where active chlorine primarily exists as HClO. Active chlorine participates in the organics degradation of organic compounds through reactions in the liquid phase (Eq. 7) or at the electrode surface ((8), (9)) [24].

  Origanics + ClO- → intermediates → CO2 + Cl- + H2O

  (7)

  MOx(•OH) + Cl- → MOx(HOCl) + e-

  (8)

  $ \begin{aligned} & \text { Origanics }+\mathrm{MO}_{\mathrm{x}}(\mathrm{HOCl}) \rightarrow \text { intermediates } \rightarrow \mathrm{MO}_{\mathrm{x}}+\mathrm{CO}_2 \\ & +\mathrm{Cl}^{-}+\mathrm{H}_2 \mathrm{O}+\mathrm{H}^{+} \end{aligned} $

  (9)

  The contribution rate of direct oxidation increases gradually as the reaction time increases (Fig. 2c, Eq. 10) [36].

  M + R → M(•R) + ne- → product

  (10)

  As the energizing time increases, the electron transfer and contribution both gradually increases. However, the contribution rate of hydroxyl free formation decreased gradually with the increase of reaction time (Fig. 2c). This is mainly because a portion of the hydroxyl group catalyzes the oxidation of the target, while another portion is converted into active chlorine (Eq. 11) [37], which also catalyzes the degradation of the target.

  •OH + Cl- → OH- + Cl•

  (11)

  With the increase of reaction time, the contribution of chlorine slightly decreased (Fig. 2c). In the oxidation process, the chemical and electrochemical generation of chlorate can occur (Eqs. 12–14) as well as the reduction at the cathode (Eqs. 12–15) [24]. This leads to a reduction in the active chlorine used for oxidation reactions. In accordance with Fig. 2b, the formation rate of active chlorine gradually slowed down as the reaction progresses.

  ClO- + 2HClO → ClO3- + 2Cl- + 2H+

  (12)

  6ClO- + 3H2O → 2ClO3- + 4Cl- + 6H+ +1.5O2 + 6e-

  (13)

  Cl- + 3H2O → ClO3- + 6H+ + 6e-

  (14)

  ClO- + H2O + 2e-→ Cl- + 2OH-

  (15)

  To sum up, the degradation of MIT may have the following situations: Firstly, MIT can be directly oxidized and degraded on the anode plate. Secondly, the ^•OH and active chlorine (HClO) generated near the anode plate can oxidize and degrade MIT indirectly. Thirdly, the Sn-Sb-Ce/GAC particle electrode directly degrades MIT after repolarization. Finally, the ^•OH and active chlorine (HClO) generated by the depolarization of the Sn-Sb-Ce/GAC particle electrode indirectly oxidizes and degrades MIT (Fig. 3).

  Figure 3

  

  Figure 3.  MIT degradation process illustration.

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  To verify the performance of the 3D electrode system, we used the secondary effluent from a sewage treatment plant, and added MIT and salt to prepare ROC wastewater.

  From Fig. 4a, the chemical oxygen demand (COD) and total organic carbon (TOC) removal rates, instantaneous current efficiency (ICE) and mineralization current efficiency (MCE) of the 3D electrode system were superior to those of the 2D electrode system. This indicated that adding Sn-Sb-Ce/GAC particle electrodes could effectively enhance the removal of organic pollutants and improve the energy utilization rate of the system. After 120 min, the COD removal rates of the 2D and 3D electrode systems were 29.4% and 42.6% respectively, while the TOC removal rates were only 10.6% and 13.7%. This suggests that the system has a certain degradation ability for organic matters in ROC wastewater, but it is difficult to completely remove pollutants by mineralization. The ICE decreased significantly during the reaction, and the MCE also showed a downward trend. The main reason is that the pollutants are continuously oxidized and degraded in the early stage of the reaction, leading to a decrease in the concentration of organic matters in the solution that can be oxidized and mineralized by the system. The accumulation of intermediate products, such as small acids and alcohols, change the physicochemical properties of the ROC wastewater, which in turn intensifies the occurrence of side reactions, thus reducing the ICE of the system.

  Figure 4

  

  Figure 4.  2D and 3D performance analysis of ROC water. (a) COD, TOC, MIT removal rate and ICE, MCE, ƞ. (b) EE/O. (c) The cell voltage. (d) UV₂₅₄. (e) Humus degradation. The 3D electrode system reaction condition [MIT] = 50 mg/L, current density = 20 mA/cm², initial pH 5, [Na₂SO₄] = 0.2 mol/L, particle electrode is 5 g/L. The 2D electrode system has the same reaction conditions as the 3D electrode except that there are no particle electrodes.

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  From Fig. 4a, there is no significant difference between the MIT removal rate and oxidation efficiency (ƞ) of 2D and 3D electrode system. After 30 min of reaction, the 3D electrode system could achieve a removal rate of 96.5% for MIT, while the 2D electrode system also reached a removal rate of 94.0% for MIT. It is because the ROC wastewater contains a large amount of Cl^- (6630 mg/L), SO₄^2- (12, 950 mg/L), which under electrocatalysis action produce chlorine free radicals, ^•OH and ^•SO₄^-, thus resulting in a higher degradation efficiency. In addition, the removal rates of MIT by both systems were better than those of COD and TOC. This also indicates that MIT is easily catalytically oxidized to form small molecular intermediates, but it is difficult to completely mineralized into carbon dioxide. Like the ICE, the ƞ of MIT also decreases over time. When the reaction proceeds for 30 min, the ƞ of the 3D electrode system drops from 34.07% to 9.42%, a total decrease of 72.4%, indicating that the energy utilization rate of the system is also gradually decreasing.

  As shown in Fig. 4b, the Electrical energy per order (EE/O) of both electrocatalytic systems was at an extremely low level, with no significant difference. However, the EE/O of the 3D electrode system was relatively low. Additionally, the EE/O of the actual water sample at any time was far lower than that of the experimental water sample prepared with pure water. Under the same conditions, the EE/O of electrocatalytic oxidation of pure water system for 20 min was 2.44 kWh/m³, which was 4.1 times that of the actual water sample. The energy consumption of the reaction system is mainly determined by the cell voltage at both ends of the electrode. Since the electrode cell voltage is directly affected by electrolyte concentration in wastewater. The total dissolved solids (TDS) of the ROC wastewater used in the experiment was 39, 400 mg/L, with a conductivity greater than 60 μS/cm. Therefore, the actual wastewater provides a natural electrolyte for the electrocatalytic oxidation system. In fact, the higher electrolyte concentration improves the electron transfer efficiency of the system, which allows the pollutants to be degraded to the maximum extent with relatively low energy consumption.

  As can be seen from Fig. 4c, the system voltage gradually increased during the reaction process, resulting in a rise in the EE/O. At the same time, compared to the 2D electrode system, the cell voltage of the 3D electrode system remained relatively stable. Under stable operation, the mass transfer efficiency of pollutants and oxidizing substances is higher, leading to improved efficiency for the system.

  From Fig. 4d, it can be observed that the UV₂₅₄ of both systems initially decreased and then increased. The initial UV₂₅₄ of the actual wastewater was 1.49, and after 20 min the UV₂₅₄ of both systems dropped rapidly to 0.502 and 0.433. After 60 min of reaction, the UV₂₅₄ began to rise gradually. At 120 min the UV₂₅₄ values were 0.89 and 0.67, respectively. In the initial stage of the reaction, both electrocatalytic systems could oxidize and remove organic compounds with unsaturated bonds from the wastewater. However, as the reaction progresses, some pollutants, including MIT, generate intermediate products that were difficult to be completely mineralized. These intermediates were absorbed at around 254 nm. The UV₂₅₄ of the system increases due to the accumulation of intermediates that are difficult to fully mineralize. In addition, it can be seen from Fig. 3d that the 3D electrode system generated relatively fewer intermediates, which further indicated that the MCE and ƞ of the 3D electrode system were superior to those of the 2D electrode system.

  Through the comparison and identification of Open Fluor database, it was identified that the ROC wastewater contained two fluorescent components: Ex/Em = 335/415, corresponding to the fluorescence response region of marine humus (MH), and Ex/Em = 367/477, corresponding to the fluorescence response region of terrestrial humus (TH). From Fig. 4e, it can be observed that with the extension of time, the maximum fluorescence intensity of both MH and TH components showed a declining trend. However, the MH content was greater than the TH content. The maximum fluorescence intensities of the two components at 120 min were 0.005 (R.U.) and 0.008 (R.U.), respectively. Hence, the results indicate that the 3D electrode system can remove and transform almost all fluorescent substances as the reaction proceeds. It is evident that the 3D electrode system has a good removal effect on fluorescent substances.

  To sum up, it is considered that the Sn-Sb-Ce/GAC of 3D electrode system possesses great application potential in the field of ROC wastewater treatment. This study provides theoretical and technical references for the efficient treatment and resource recovery of typical industrial high-salinity organic wastewater, such as those from reverse osmosis processes. This holds significant practical importance for improving water resource utilization and protecting aquatic ecosystems.

  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

  Feng Xu: Writing – review & editing, Writing – original draft, Visualization, Methodology, Investigation, Formal analysis, Data curation, Conceptualization. Yuqiu Liu: Methodology, Data curation, Conceptualization. Shujiao Xu: Resources, Investigation. Jinxin Zhang: Visualization, Investigation. Lei Liao: Investigation. Jiguang Guo: Validation, Investigation. Weiyu Jiang: Visualization, Investigation. Hongzhe Dong: Investigation, Conceptualization. Qinxue Wen: Writing – review & editing, Conceptualization. Zhiqiang Chen: Writing – review & editing, Supervision, Conceptualization.

  Acknowledgments

  The authors appreciate the financial supports from Major Science and Technology project of China Power Engineering Consulting Group Co., Ltd. "Research on Green and digital Intelligent Technology of Sewage Treatment Plant" (No. CEEC2023-ZDYF-09) and Technology Innovation Ability Improvement Project of Shandong Province, China (No. 2022TSGC1247).

  Supplementary materials

  Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.cclet.2025.111332.

  
  
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  Figure 1  (a) SEM of GAC (10 K). (b) SEM of Sn-Sb-Ce/GAC (10 K). (c) SEM of Sn-Sb-Ce/GAC (20 K). (d) EDS of Sn-Sb-Ce/GAC particle electrode. (e) XRD pattern of particle electrode.

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  Figure 2  (a) The effect of TBA and MeOH on the removal effect of MIT and the curve of degradation kinetics. (b) The amount of active chlorine produced during electrocatalytic oxidation. (c) Contribution rate of different oxidation modes in the electrocatalytic oxidation of MIT by 3D electrode system. The 3D electrode system reaction condition [MIT] = 50 mg/L, current density = 20 mA/cm², initial pH 5, [Na₂SO₄] = 0.2 mol/L, Sn-Sb-Ce/GAC particle electrode was 5 g/L.

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  Figure 3  MIT degradation process illustration.

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  Figure 4  2D and 3D performance analysis of ROC water. (a) COD, TOC, MIT removal rate and ICE, MCE, ƞ. (b) EE/O. (c) The cell voltage. (d) UV₂₅₄. (e) Humus degradation. The 3D electrode system reaction condition [MIT] = 50 mg/L, current density = 20 mA/cm², initial pH 5, [Na₂SO₄] = 0.2 mol/L, particle electrode is 5 g/L. The 2D electrode system has the same reaction conditions as the 3D electrode except that there are no particle electrodes.

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