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Review

Electrochemically Assisted Persulfate Oxidation of Organic Pollutants in Aqueous Solution: Influences, Mechanisms and Feasibility

by
Jianting Sun
1,
Wei Zheng
1,
Gang Hu
1,
Fan Liu
1,
Siyuan Liu
1,
Lie Yang
2,* and
Zulin Zhang
2,3,*
1
Wuhan Huantou Qianzishan Environmental Industry Co., Ltd., Wuhan 430000, China
2
School of Resources and Environmental Engineering, Wuhan University of Technology, Wuhan 430070, China
3
The James Hutton Institute, Craigiebuckler, Aberdeen ABI5 8QH, UK
*
Authors to whom correspondence should be addressed.
Catalysts 2023, 13(1), 135; https://doi.org/10.3390/catal13010135
Submission received: 10 December 2022 / Revised: 29 December 2022 / Accepted: 5 January 2023 / Published: 6 January 2023
(This article belongs to the Special Issue Advanced Catalytic Material for Water Treatment)
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Abstract

:
Electrochemically (EC) assisted persulfate (PS) oxidation processes (EPOPs) have gained increasing attention in recent years. In this review, the current status and prospects of EC/PS degradation of organic pollutants are discussed and summarized. It was found that the oxidation of most organic contaminants could be significantly enhanced or accelerated using the combination of EC and PS compared to single treatments. Moreover, the effects of various operational variables on the removal of organic contaminants were investigated. Some variables are highly sensitive, and the optimal conditions are case-specific. Regarding the degradation mechanisms, radical-induced reactions and nonradical reactions both exist for the elimination of organic contaminants. Oxidants (including S2O82− and SO4•−) can be produced from SO42− near the anode, which is a unique feature of EPOPs. In some studies, the electrical energy consumption of EPOPs has been controlled to a reasonably low level in lab-scale attempts. Although there are still a few drawbacks or difficulties (e.g., potential electrode fouling, dependency on batch mode) for large-scale applications, EPOPs offer a promising alternative to traditional advanced oxidation techniques.

1. Introduction

Hydrogen peroxide (H2O2) is a typical oxidant in traditional advanced oxidation processes. In recent years, persulfate (PS, including permonosulfate (PMS) and peroxodisulfate (PDS)) has gained great interest as a promising alternative to H2O2 for advanced oxidation processes due to its high redox potential and longer lifetime [1,2,3]. Although the price of PS (e.g., ammonium persulfate, USD 715–USD 815 per ton) is higher than that of H2O2 (USD 470–USD 610 per ton) (https://www.alibaba.com/, accessed on 1 November 2022), PS can be easily stored and delivered in a solid form due to its stability in the inactive state [1,4]. In addition, the chemical cost was reported to be much lower than the energy cost for the treatment of some refractory organic pollutants (e.g., per- and polyfluoroalkyl substances) [5]. In view of standard electrode potentials, SO4•− (E0 = 2.5V~3.1 V) can be formed when the O-O bonds in PS (E0 = 2.1 V) are broken using various activation methods [2,6]. Traditional •OH radicals have a standard reduction potential of 2.7 V in an acidic solution and 1.8 V in a neutral solution [7]. Several activation methods have been explored in the past decade, including energy input (e.g., heat [8,9], ultrasound [2], ultraviolet rays [10], electrochemistry [11], and microwave [12]), nano-materials [13], transition metals [14,15], and carbon materials [16]. Among these techniques, electrochemical activation is an emerging choice and has attracted increasing attention during the past 12 years (Figure 1) based on statistical data from the Web of Science (www.webofscience.com, accessed on 1 November 2022). Electrochemical oxidation technology can be utilized alone for the decomposition of organic pollutants, e.g., ciprofloxacin [17], enrofloxacin [18], and perfluorooctane sulfonate [19].
It was reported that at various Fe2+ dosages (2 mg/L, 4 mg/L, 6 mg/L and 8 mg/L), the phenol degradation and COD removal observed were much higher in electro-Fenton processes than in traditional Fenton processes [20]. In particular, electro-Fenton processes could be significantly enhanced when PS was added [21]. The addition of PS can provide sufficient oxidants and facilitate the degradation of organic contaminants [22]. The combination between electrochemistry (EC) and PS is a promising substitution for traditional advanced oxidation processes. On the one hand, electron transfer from the electrochemical process can activate PS to generate SO4•− [23]. The values of the standard one-electron reduction potentials of SO4•− and S2O82− are 2.437 +/− 0.019 and 1.44 +/− 0.08 V vs. the standard hydrogen electrode (SHE), respectively [24]. SO4•− could attack target compounds and generate byproducts including SO42− [8,9]. On the other hand, S2O82− can be generated from SO42− on the surface of an anode [25,26,27]. The regeneration process of S2O82− can provide an effective supply of oxidants. Compared to other PS activation methods, this unique feature of electro-activated PS is beneficial for cost reduction in further full-scale applications, which could explain why the electrochemical activation method has gained increasing interest.
A number of attempts have been made to eliminate various organic pollutants using EC-assisted PS-based oxidation processes (EPOPs), including dyes [23,28], pharmaceuticals [29,30], and herbicides [31], among others. The performance of EPOPs in these studies was not the same, mainly due to the different operational variables and degradation mechanisms. Thus, a detailed review is necessary to outline and discuss the current status and trends in EPOPs. Although several reviews on electrochemical advanced oxidation processes (EAOPs) were reported in recent years [32,33,34,35], these reviews mainly focused on H2O2-based processes. Li et al. [36] conducted a mini-review involving EAOPs–PS systems and EAOPs–PMS systems, which primarily discussed the reaction mechanisms of homogeneous (Fe2+ and Fe3+) and heterogeneous catalysts (Fe, Mn, and Co oxides) [36]. Nevertheless, the radical and non-radical mechanisms of organic pollutant degradation in EPOPs have not been systematically summarized yet. The conversion between S2O82− and SO42− was not discussed in the previous review. In addition, the synergistic influences of EC and PS and operational variables in EPOPs were not reviewed yet. Therefore, this review places emphasis on the following targets: (1) to summarize the synergistic influences of EC and PS on the removal of organic contaminants; (2) to discuss the influences of operational variables on organic contaminant removal; (3) to analyze and outline the predominant degradation mechanisms of EPOPs; and (4) to investigate the energy consumption of EPOPs.

2. Synergistic Influences of EC and PS

Electrochemical oxidation technology can be utilized alone for the decomposition of organic pollutants [17,18,19]. Electrochemical oxidation of contaminants occurs through two reaction paths including direct anodic oxidation and indirect oxidation [37,38]. Both mechanisms involve electron transfer, which can be used to activate PS [23]. In view of the promising combination of EC and PS, the performances of single and combined EC and PS treatments are compared to analyze their synergistic effects on the decomposition of various pollutants (Table 1). It is worth mentioning that significant differences are observed in the outcome of these methods for the removal of various organic contaminants, probably due to the differences in operational variables and molecular structures. The degradation efficiencies of the “EC + PS” combination are better than those of single EC or PS oxidation processes for most of the selected target compounds. Moreover, single EC is more efficient for the majority of selected organic pollutants than sole PS treatment. Unactivated PS was ineffective during the degradation processes probably due to its lower redox potential (S2O82−, E0 = 2.1 V) than that of SO4•− (E0 = 2.5 V~3.1 V) [23,28,29]. Single electrochemical oxidation was also found to be inefficient for several target compounds, e.g., bisphenol A [29] and carbamazepine [39] with 3.5% and 2.18% removal efficiencies, respectively. Once the two methods are combined, significant enhancement can be achieved for bisphenol A and carbamazepine with 76.5% and 97.76% removal efficiencies, respectively. It was found that the concentrations of diuron remained unchanged when treated with PS alone, and only 15% of diuron was removed with EC treatment [40]. Nevertheless, over 77% of diuron was eliminated when the EC/PS system was employed [40]. Analogously, negligible oxcarbazepine (OXC) was degraded using PS alone and a similar low removal was obtained with the application of electrocoagulation alone [30]. In contrast, a high removal rate of ~75% was achieved within only 10 min when 0.5 mM PS was added [30]. On the whole, the combination of EC and PS is a promising advanced oxidation system for aqueous refractory organic contaminants and trace contaminants.

3. Influence of Operational Parameters on the Degradation Efficiency of EPOPs

3.1. Electrode Material

The electrode material is one of the most vital factors determining the efficiency and cost of the electrochemical process [58]. The electrocatalytic activity, service life, and stability are the most important characteristics of the anode/cathode materials [58]. There has been increasing concern for electrode materials in EPOPs [36]. The commonly used electrodes include anode materials such as mixed metal oxides (e.g., Ti/RuO2-IrO2) [23,29,59], iron sheet [42,60,61], carbon felt [62], cathode with titanium [63], stainless steel [55], and carbon fiber [64]. Some electrode materials do not react with electrolytes, while several materials (e.g., iron) participate in the redox reactions in EPOPs [60]. It is known that Fe2+ is a typical activator of PDS and PMS via Equations (1) and (2) [36,65]. The generated Fe3+ can be reduced to Fe2+ by the cathodic reaction (Equation (3)) [66], and then Fe2+ can activate PS in the solution again. Therefore, Fe2+ has been frequently applied to enhance the degradation efficiency of EPOPs due to its outstanding ability to activate PS [23,55,67,68]. However, excessive Fe2+ can simultaneously scavenge sulfate radicals via Equation (4) [66], which would undoubtedly reduce the oxidation efficiency. In particular, the iron electrode could be used to control the release of Fe2+ via the anodic reaction (Equation (5)) to optimize the degradation process, as reported in previous studies [66,69]. Moreover, the effective current transfer between the anode and cathode would probably be reduced by the formation of a hydrated iron hydroxide/oxide film on the surface of the iron anode. This problem could be avoided by using an acidic pH condition. The positive effect of acidic pH will be discussed in Section 3.4 in more detail.
Fe2+ + S2O82−→Fe3+ + SO42− + SO4•−
Fe2+ + HSO5−→Fe3+ + OH− + SO4•−
Fe3+ + e−→Fe2+
Fe2+ + SO4•−→Fe3+ + SO42−
Fe(s)→Fe2+ + 2e−
In addition to the iron electrode, there are also significant differences in the degradation efficiency among non-active electrodes. It was found that ampicillin removal rates after 10 min of reaction were 68% and 39% for boron-doped diamond (BDD) and platinum anodes, respectively [70]. The strong contact of hydroxyl radicals on the surface and the lower potential window of platinum were considered to be the main reasons for the observed phenomenon [70]. Furthermore, the oxygen evolution overpotential of electrode materials is believed to be vital for the oxidation of organic contaminants [37]. Therefore, high O2 overvoltage anodes such as BDD are usually preferred [37]. Cathode materials are also important for analyte removal in EPOPs. The removal of ciprofloxacin and total organic carbon was compared between BDD/Pt and BDD/Gr electrode pairs in a single-cell rotating disk electrode system to evaluate how cathode materials affect EPOPs during analyte removal [71]. The results indicated that ciprofloxacin and total organic carbon (TOC) degradation rates were statistically insignificant with Pt- and Gr-cathodes, and graphite was a suitable substitute for platinum [71]. During the PS activation process, the concentration of sulfate ions would inevitably increase. Some of the sulfate ions could be converted to PS again [25]. Nonetheless, the final concentration of sulfate ions in the effluent was not investigated in previous studies.

3.2. Electrolyte Type and Concentration

An electrolyte is the medium of electrochemical reactions, which can provide the precursors of reactive oxidizing species [71,72]. A suitable electrolyte is necessary for the efficient removal of organic pollutants in EPOPs. The effects of nitrate, sulfate, and persulfate electrolytes on ciprofloxacin removal in an electrochemical reactor were compared [71]. Ciprofloxacin was efficiently removed with EC/PS treatment, achieving 84% after 120 min and reaching 90% after 240 min [71]. The removal rate in sulfate was slower, although a similar degree of removal was achieved within 240 min [71]. Nevertheless, a much slower removal of ciprofloxacin was observed when nitrate was employed as the electrolyte, reaching 90% in 24 h [71]. Cathodic persulfate activation occurred when persulfate was added as the electrolyte, and SO4•− could attack ciprofloxacin effectively [71]. Moreover, S2O82− could be generated from SO42− via Equation (6) [26,27,73,74]. The electrochemical oxidation of multiple organic compounds was also studied while varying the background electrolyte among NaCl, Na2SO4, and NaClO4 [75]. When the electrolyte exchanged from ClO4− to Cl− or SO42−, the anodic oxidation performance became substrate-specific, probably because the electrolyte-derived reactive species (Cl2•−, SO4•−) contributed to the overall electrochemical oxidation efficiency [75]. Overall, Cl− and SO42− acted as the precursors of reactive oxidizing species (Cl2•−, SO4•−) and contributed to the organic decomposition in EPOPs, while ClO4− and NO3− could only depend on anode oxidation.
2SO42−→S2O82− + 2e−
Apart from the electrolyte type, the electrolyte concentration is another major concern for the medium optimization. There are several studies focusing on the influence of electrolyte concentration on the degradation of organic contaminants [29,39,74]. It is well known that conductivity increases with an increase in electrolyte concentration. Nevertheless, the relationship between electrolyte concentration and the degradation of organic contaminants is much more complicated. As previously mentioned, ClO4− and NO3− can only depend on anode oxidation via •OH reactions. It was observed that the reaction rates marginally improved over the range of 0.75 h−1 to 1.06 h−1 with 8.5–60 mM NaNO3 [74]. Regarding Na2SO4, it was found that the decomposition rate of bisphenol A decreased when the concentration increased from 50 to 200 mM [29], probably due to the voltage decline of the system [39]. A similar phenomenon was observed in the degradation of carbamazepine in EPOPs. It is worth mentioning that a sufficient amount of SO42− is necessary for efficient oxidation in EPOPs. The concentration of 40 mM was determined to be optimal when the Na2SO4 concentration was varied from 5 mM to 40 mM [74]. Therefore, electrochemical oxidation is substrate-specific when an electrolyte is the precursor of reactive oxidizing species, and an appropriate electrolyte concentration is vital for the efficient degradation of organic contaminants.

3.3. Current Density

Current density (CD) has been frequently investigated for optimizing the electrochemical oxidation processes [23,28,29]. Previous studies involving CD optimization are presented in Table 2. On the one hand, a higher current density improved the generation of sulfate radicals via an electron transfer reaction [29]. On the other hand, increasing the CD led to faster ferrous ion production and regeneration rate when an iron anode, Fe2+, and Fe3+ were used [29,39]. In addition, S2O82− can be generated from SO42− in the electrolyte via Equation (6), as mentioned above [25]. It was found that a significant increase in the production of S2O82− was achieved at 30 mA·cm−2 and 60 mA·cm−2, compared to that at 15 mA·cm−2 [25]. This process could be considered as the in situ generation of oxidant supply. Nevertheless, it was also observed that there was no enhancement in the pollutant degradation when further increasing the CD after it reached a limit [39,76]. The limit depended on the target contaminants. When the current density reached the limit, the inhibition effect on pollutant removal became significant. Han et al. [39] found that the removal rate of carbamazepine was increased as the CD was switched from 3.57 to 7.14 mA/cm2. Nonetheless, no significant improvement was observed when the CD increased from 7.14 to 14.28 mA/cm2 [39]. Similarly, the COD removal from landfill leachate using oxidation, as well as the total COD removal, increased significantly as the CD increased from 40 mA to 80 mA, while a further increase in the current did not lead to a clear improvement in COD removal [76]. This phenomenon could be attributed to some side reactions such as hydrogen evolution (Equation (7)) and PDS reduction (Equation (8)). In addition, it was noticed that the optimal current densities of some compounds (e.g., phenol and ciprofloxacin) reported in different publications were significantly different. For most entries, higher current densities could achieve higher removal rates of organic pollutants. The main reason for this phenomenon was the different range of the selected current density. These target pollutants were degraded with higher removal rates under higher current densities.
2H+ + 2e−→H2
S2O82− + 2e−→2SO42−
Inhibition effects on the removal of organic pollutants were found with further increases in the CD [30,53,81]. The removal rate of oxcarbazepine decreased rapidly when the CD increased over 8.3 mA/cm2, which can be attributed to the quenching of •OH and SO4•− due to the excess Fe2+ produced from the anode [30]. Analogously, the ciprofloxacin removal rates at the contact time of 40 min and a CD of 0.94, 1.55, 1.75, 2.75, and 4.25 mA/cm2 were 79.65, 87.55, 89.32, and 80.31%, respectively [53]. It was obvious that the CIP removal efficiency decreased with the increase in CD to 4.25 mA/cm2 [53]. A similar phenomenon was observed in a study of Basic Red 18 decolorization using EPOPs [81]. On the whole, increasing the current density could promote the degradation of organic contaminants within an appropriate limit. Further increases would inhibit the oxidation processes probably due to hydrogen evolution, PDS reduction, and excess Fe2+ generation.

3.4. Initial pH

The initial pH plays an important role in the degradation of contaminants in SO4•− -grounded systems [2,91]. The influences of pH on the removal efficiencies in EPOPs have been frequently investigated (Table 3). Most of the previous studies found that an acidic pH was favorable for the degradation of selected organic contaminants in EPOPs. The phenomenon was attributed to the notable decline in soluble oxygen gas in wastewater at low pH values, which could facilitate the contact between persulfate anions and the cathode and result in the formation of sulfate radicals [43,44]. Moreover, when the reaction occurred in the presence of Fe2+ or Fe3+, a Fe-containing precipitate was generated to retard the oxidation process [29]. When the pH was greater than 4.0, ferric ions formed some Fe3+ oxyhydroxides, such as FeOH2+, Fe2(OH)24+, Fe(OH)2+, and Fe(OH)3 (Equations (9) and (10)) [29,92]. A similar phenomenon was observed in the presence of Fe2+ via Equation (11) [29,39]. The precipitate formation is believed to inhibit the activation reaction between Fe2+ and PDS [29]. In addition, SO4•− can react with OH- to generate •OH (a weaker oxidant) via Equation (12) under alkaline conditions [2].
Fe3+ + H2O → FeOH2+ + H+
2Fe3+ + 2H2O →Fe2(OH)24+ + 2H+
Fe2+ + H2O →FeOH+ + H+
SO4•− + OH−→SO42− + HO•
Nonetheless, a low pH is not always favorable for the oxidation of organic contaminants. When the initial pH of Orange Ⅱ decreased from 3 to 2, the initial decolorization rate and efficiency decreased slightly [28]. Similarly, a further decrease in pH from 4.00 to 2.00 caused a significant drop in the degradation of tetracycline hydrochloride [57]. A high H+ concentration would promote the side reaction of H2 generation via Equation (7). In particular, it was found that the rate constant was significantly increased under acidic (pH = 3) and alkaline conditions (pH = 12) [83]. Therefore, it could be deduced that alkaline activation plays a vital role in the initial stage when the initial pH is set at 12. In addition, pHPZC of the applied catalysts could be relevant to optimal pH when a catalyst is combined with the EC/PS processes [42,51,80]. Moreover, a pH adjustment for efficient removal contributes to the operational cost of EPOPs. Overall, the optimal pH in EPOPs depends on a series of factors including the soluble oxygen, Fe2+/Fe3+, and catalyst.

3.5. PS Concentration

The effects of the initial PS concentration in EPOPs are summarized in Table 4. As demonstrated in Table 4, the removal rates exhibited a significant increasing trend with the increase in PS concentration in some previous studies [23,28,87]. This phenomenon could be attributed to the significant increase in the number of sulfate radicals [54]. PS is the chief source of sulfate radicals in the system, and more reactive radicals would be generated to degrade the organic contaminants at a higher PS concentration [23]. Some other studies found that when the PS concentration reached a certain limit, a further increase in PS concentration did not promote and even inhibited the degradation efficiency [29,51,85]. It was observed that the TOC removal efficiency exhibited an increasing trend with the addition of persulfate anions in the prepared aniline solution [44]. However, in the presence of excess persulfate anions (3.0 wt%), the removal rate of aniline was decreased [44]. Similar phenomena were found in other studies involving hexachlorocyclohexanes [84], methyl orange [69], sodium dodecylbenzene sulfonate [68], and dinitrotoluenes [96]. The side reactions between S2O82− and reactive radicals (Equations (13) and (14)) could probably explain this phenomenon when excess PS is added, which would consume more S2O82− [29,85]. It is evident that an excess persulfate anion concentration would inhibit the formation of hydrogen peroxide from sulfate radicals [43]. Additionally, the recombination of sulfate radicals (Equation (15)) may also occur under excess PS concentrations [51]. The mentioned side reactions are consistent with other PS-based processes [2]. In particular, it was noted that the optimal PS concentrations of some compounds (e.g., aniline and atrazine) reported in different publications were significantly different. Indeed, the results were consistent with the above analysis. Three atrazine-based studies indicated that the removal rates increased with the increase in PS concentration. When the range of PS concentrations was extended, excessive PS concentration caused side reactions, thus decreasing the removal efficiency. The limit values are target-dependent and easily affected by the operational conditions. Therefore, the selection of PS concentration should consider the degradation rates and economic costs.
S2O82− + SO4•−→S2O8•− + SO42−
S2O82− + •OH→S2O8•− + OH−
SO4•− + SO4•−→S2O82−

3.6. Temperature

Temperature is an important parameter for the solution viscosity [99] and mass transfer coefficients [100]. As a result, temperature is frequently considered due to its crucial influence on the degradation of organic contaminants in EPOPs [43,44]. It was observed that the degradation efficiency of TOC (aniline solution) at 318 K was higher than that at 313 K during the electro-activated persulfate process, indicating that a higher reaction temperature was beneficial to the mineralization of aniline [43]. A similar phenomenon was also found in other studies using EPOPs, including the treatments of Disperse Blue 3 [83], carbamazepine [39,51], phenol [61], landfill leachate [93], and petrochemical wastewater [94]. The enhancement of the degradation efficiency with an increase in temperature could be ascribed to multiple mechanisms. Firstly, an increase in temperature is beneficial for mass transfer in the reaction system [61], which would accelerate the reaction process. Secondly, heat can activate PS, promoting the generation of SO4•− [101]. Finally, the amount of oxygen gas dissolved in wastewater would decrease at higher reaction temperatures, which could facilitate the contact between the persulfate anions and cathode and lead to the formation of sulfate radicals [43,44]. Nevertheless, it was still found that a continued increase in temperature did not promote the removal of organic contaminants in several studies [76,79]. The negative influence of extremely high temperatures could be attributed to some quenching reactions (Equations (13) and (15)) [76].

3.7. Nitrogen/Oxygen Dosage

Nitrogen and oxygen dosages also affect the generation of sulfate radicals during the electro-activated persulfate process [43,44,54]. The influence of the flow rate of nitrogen and oxygen on the electrolytic behavior has been investigated in the degradation of dinitrotoluenes, aniline [43,44], and Disperse Blue 3 [83]. It was found that the TOC removal efficiency of industrial wastewater containing dinitrotoluenes increased with the increase in nitrogen dosage [54]. In contrast, the degradation efficiency of TOC decreased upon raising the oxygen flow rate [54]. Similarly, under O2-saturated conditions, a considerable amount of H2O2 was generated, leading to lower PS decomposition and 25% less decolorization of Disperse Blue 3 compared to the control reaction [83]. However, under N2-saturated conditions, an effective decolorization of about 99% was achieved [83]. The phenomenon may be attributed to nitrogen enhancement in the amounts of sulfate radicals generated via the cathodic reduction of persulfate anions, which would compete with oxygen for electrons (Equations (16) and (17)) [44]. Upon introduction of nitrogen gas, the oxygen formed from anodic oxidation of water would be expelled from the wastewater, resulting in the production of a lower amount of H2O2 [54]. Conversely, there was an increase in the number of dissolved oxygen molecules, which intensely competed with persulfate anions for electrons at the cathode and then inhibited the PS activation process [43].
S2O82− + e−→SO4•− + SO42−
O2 + 2H+ + 2e−→H2O2

3.8. Coexisting Ions and Natural Organic Matter

The effects of coexisting inorganic anions on the degradation efficiency of EPOPs have been well studied due to their ubiquitous distribution in various types of water [40,41,50]. The effects of Cl−, HCO3−, CO32−, NO3−, SO42−, and H2PO4− have been mentioned in previous studies [31,40,50]. Generally, these inorganic anions have negative effects on sulfate and hydroxyl radicals because these ions can quench the sulfate and hydroxyl radicals via Equations (18)–(27) [31,40,50]. Less-reactive inorganic radical species are generated after these reactions, causing lower degradation efficiencies of the organic contaminants. Nevertheless, Cl- was found to promote the degradation of target organic contaminants in some studies using EPOPs [31,41,50,80]. Ma et al. [41] found that low concentrations of Cl- inhibited the degradation of 2,4-dinitrotoluene in the earlier part of the reaction, whereas high concentrations of Cl− facilitated the degradation of 2,4-dinitrotoluene. Similarly, it was reported that Cl− enhanced the decomposition of carbamazepine in electrochemically activated PS oxidation [50]. The influence of Cl- on degradation might change along with the structure of the pollutants [11]. If chlorine radicals (e.g., Cl• and Cl2•−) are effective for organic contaminants with certain molecular structures, then the enhanced generation of chlorine radicals (e.g., Cl• and Cl2•−) using electrochemistry could promote the degradation efficiencies in these studies. It was reported that the presence of Cl− accelerated the degradation of ATZ, with the rate constant increasing from 0.050 to 0.069 min−1. As for SO42−, it is the byproduct of PS activation reactions. Although it cannot be easily oxidized, S2O82− can be generated from SO42− via Equation (6), which could be considered as an effective supply of PS [25]. Meanwhile, sulfate ions could act as an electrolyte to improve the reactions.
SO4•− + Cl−→SO42− + Cl•
•OH + Cl−→OH− + Cl•
Cl− + Cl• →Cl2•−
SO4•− + HCO3−→SO42− + H+ + CO3•−
SO4•− + CO32−→SO42− + CO3•−
•OH + HCO3−→H2O + CO3•−
•OH + CO32−→OH− + CO3•−
SO4•− + NO3−→SO42− + NO3•
•OH + NO3−→NO3• + OH−
SO4•− + H2PO4−→H2PO4• + SO42−
Natural organic matter (NOM) is able to scavenge radicals, which can lower the degradation efficiency [50]. It is necessary to investigate the effects of NOM on the degradation of target compounds in EPOPs. Suwannee River natural organic matter [30] and humic acid [50,70] have been used as representative NOM to simulate the coexisting process. Bu et al. [30] found that the degradation rate of oxcarbazepine decreased from 0.14 min−1 to 0.05 min−1 when the dosage of NOM increased from 0 mg/L to 10 mg/L. The rapidly decreasing pseudo-first-order rate constant of oxcarbazepine degradation at low dosages of NOM can be ascribed to the sharp consumption of •OH by NOM based on the reaction rate constants [30]. Similarly, the addition of humic acid during electrochemically activated PS oxidation significantly inhibited the decomposition of carbamazepine [50]. The phenomenon could be ascribed to the reasons involved in competitive contact and oxidation [50]. Therefore, it can be deduced that competitive contact and oxidation are the predominant mechanisms of NOM inhibition.

4. Degradation Mechanisms

The removal mechanisms of EPOPs are mainly considered to be the joint contributions of the radical-induced reactions [82] and non-radical oxidation [50] (Figure 2). Radical-induced reactions, typically involving SO4•− and •OH, have been frequently reported [82,83]. The predominant radical reactions are listed in Table 5. Around the cathode, SO4•− can be generated from S2O82− via a cathodic activation, which is the chief source of SO4•− in EPOPs [82]. Under O2-saturated conditions, a considerable amount of H2O2 was generated via Equation (17), leading to a lower PS decomposition [83]. It was found that the number of oxygen molecules dissolved in the electrolyte increased, which intensely competed with persulfate anions for electrons at the cathode and then inhibited the PS activation process [43]. Although the generated H2O2 could be converted to •OH, its standard redox potential is lower than that of SO4•− [2]. Around the anode, a certain amount of hydroxyl radicals (M(•OH)) are formed as an intermediate of water discharge on the anode (M) [102], which is the original reason for anode oxidation. In addition, oxidants can be generated from SO42−, as reported in some previous studies [25,27,71]. On the one hand, S2O82− can be generated from SO42− via Equation (6), which could be considered as an effective supply of PS [25]. Other activation methods could not regenerate PS based on a literature analysis [2]. On the other hand, it was found that SO4•− could be formed anodically from SO42− using BDD electrodes [71]. These results indicate that anodic reactions also contribute to the degradation process, probably due to the oxidant and radical supply. It is worth mentioning that some reactions among radicals occur in EPOPs. For example, SO4•− can be converted to •OH via the reaction with OH−/H2O [51]. Additionally, the recombination of SO4•− and •OH occurs when the radical concentrations reach a certain level [51,103].
In addition to radical oxidation, non-radical oxidation also contributes to the removal of contaminants [45,49,50,56]. It is known that methanol can scavenge both •OH and SO4•−, and tert-butanol can be used to scavenge •OH [50,104]. Methanol (MeOH) was commonly employed as the scavenger of SO4•– (k = 2.5 × 107 M−1 s−1) and •OH (k = 9.7 × 108 M−1s−1), while tert-butyl alcohol (TBA) was applied as the scavenger for •OH (k = 6.0 × 108 M−1s−1) [105]. Surprisingly, it was observed that the addition of excess methanol or tert-butanol at a dose up to 200 times the PDS concentration during the electrochemically activated PDS process had negligible impact on the degradation of carbamazepine [50]. Similar results were reported in other studies involving the degradation of sulfamethoxazole [56] and aniline [45]. NaN3 is usually employed for 1O2 quenching [105]. Moreover, Nie et al. [45] found that the addition of NaN3 had negligible impact on the degradation of acyclovir, indicating that singlet oxygen was not responsible for the EC activation of PDS at a multi-walled carbon nanotube (MWCNT) cathode. The above results indicate that nonradical oxidation mainly contributed to the degradation of target organic compounds. It is known that electrical discharge in an electric double layer induces the stable PDS molecule into a transition (activated) state and that the activated PDS molecule has high reactivity, which can degrade organic compounds [50]. Analogously, Nie et al. [45] suggested that the MWCNT-adsorbed S2O82− with a modified electronic structure was highly reactive towards organic compounds. Although suitable in situ characterization methods for investigating the modified electronic structure of PDS are lacking at present [45], this explanation provides a plausible non-radical mechanism. Until now, the non-radical mechanisms have not been adequately studied. More in-depth investigations are still needed to reveal the exact non-radical mechanisms.
It is worth mentioning that various other methods have been combined with EPOPs, as demonstrated in Table 1. These methods were shown to be effective at enhancing the removal efficiencies using EPOPs (Table 1). Ultrasound and Fe2+/Fe3+ are selected for discussion in this study. Ultrasound enhancement could be attributed to cavitation, PS activation, and mass transfer intensification [2]. The decolorization efficiency of Acid Orange 7 increased with the increase in Fe2+ concentration due to the PS activation of the available Fe2+ [23]. Therefore, more SO4•− are generated to oxidate bisphenol A with Fe2+ involvement [29]. In addition, some heterogeneous catalysts (e.g., Fe/SBA-15 [85], MnO2 [42]) were used to promote the removal efficiencies. These combination methods also play important roles in the degradation of organic contaminants using EPOPS.

5. Energy Consumption and Limitations

5.1. Energy Consumption

Energy consumption is a key point in evaluating a method of organic contaminant treatment, especially for field applications [76]. Although no cost comparison has been made among various PS activation methods, there are several studies involving the energy consumption of EPOPs in the past decade (Table 6). It can be seen that the electrical energy consumption of EPOPs has been controlled at a reasonably low range, which is undoubtedly favorable for further field applications. Cai et al. [85] found that the electrical energy consumption per order of magnitude (EE/O) of the EC/Fe/SBA-15/PDS process was 9.87 kWh/m3, which was much lower compared to UV/ZnO/H2O2 (172 kWh/m3) [103] and UV/[TiO2 + ZnO]/PDS (18.08 kWh/m3) [106]. Similarly, the removal rate of Acid Orange 7 reached 98.1% at 30 min in the electro/MnO2/PDS system; the EE/O of the oxidation system was 7.84 kWh/m3, which was much smaller than some photocatalysis processes [42]. In addition, the EE/O of sulfamethoxazole degradation efficiency using the EC/PDS process was calculated to be only 0.04 kWh/m3 under optimal conditions [89]. It was reported that the EE/O value for sulfamethoxazole removal was 0.46 kWh/m3 for an electron beam, 27.53 kWh/m3 for ozone oxidation, 1.50 kWh/m3 for UV-C photodegradation, and 13.2–21.6 kWh/m3 for electrochemical degradation using a Ti/SnO2-Sb/Ce-PbO2 anode [89,107,108]. Apart from the comparison with other methods, the electrode material [39] and catalyst assistance [78] are also crucial for cost reduction in EPOPs. When a Pt cathode was replaced with activated carbon fiber, the EE/O value decreased from 0.6056 kWh/m3 to 0.0788 kWh/m3 in the EC/Fe3+/PDS system [39]. It was reported that the electrical energy consumption decreased from 1.151 kWh/g to 0.744 kWh/g with the addition of a nano-Fe@NdFeB/AC catalyst [78]. Overall, electrolysis assisted persulfate oxidation is an energy-saving method for removing some organic contaminants with the appropriate operational conditions.

5.2. Limitations

There are still several limitations for the further application of EPOPs. To the best of our knowledge, the following drawbacks should be noted and considered in the future:
(1)
Most EPOPs perform well in acidic conditions. A pH adjustment for efficient pollutant removal can contribute to the operational cost of EPOPs.
(2)
A lower performance in aerated conditions is usually found in practical applications.
(3)
Potential electrode fouling needs to be eliminated in time. The electrode performance should be effectively monitored.
(4)
The concentration of sulfate ions inevitably increases during PS activation processes. The introduction of sulfate ions into the effluent should be considered later.
(5)
The formation of hydrated iron hydroxide/oxide film on the surface of an iron anode reduces the effective current transfer between the anode and cathode. Acidic reaction conditions could avoid the occurrence of this phenomenon.

6. Summary and Conclusions

The combination between electrochemistry and persulfate as electrochemically assisted persulfate oxidation processes (EPOPs) is a promising substitution for traditional advanced oxidation processes. In this review, the synergistic effects of electrochemistry and persulfate, the operational variables influencing the removal efficiency, the predominant degradation mechanisms, and energy consumption were discussed and summarized. The effects of operational variables on the removal efficiency, including the electrode material, electrolyte type and concentration, current density, pH, PS concentration, temperature, nitrogen/oxygen dosage, coexisting ions, and natural organic matter, were discussed to provide beneficial references for further research. A comparison among these variables revealed that EPOPs are complicated and target depended. In addition, it was found that the decomposition mechanisms of EPOPs mainly involved the combined contributions of radical-induced reactions and non-radical oxidation. SO4•− and •OH are the dominant reactive radicals in the radical-induced reactions. The regeneration of oxidants (including S2O82− and SO4•−) around the anode is favorable for the economic feasibility of EPOPs. Non-radical oxidation also contributes to the degradation of target compounds, while the mechanism should be further explored. The transition (activated) state of PS has high reactivity and is believed to be responsible for non-radical oxidation in EPOPs. The energy consumption estimation indicates that electrolysis-assisted persulfate oxidation is an energy-saving method for removing some organic contaminants with the appropriate operational conditions. Nonetheless, there are still several gaps in our knowledge and limitations that require further exploration, including a lack of in situ applications, lower performance in aerated conditions, unclear mechanisms of non-radical oxidation, and few toxicity evaluations of intermediate products. In the future, more actual applications of EPOPs need to be conducted for complicated real waters with trace concentrations of organic contaminants.

Author Contributions

Conceptualization, J.S. and L.Y.; methodology, J.S.; formal analysis, W.Z. and G.H.; investigation, F.L. and S.L.; data curation, W.Z.; writing—original draft preparation, J.S.; writing—review and editing, L.Y. and Z.Z.; supervision, L.Y.; project administration, L.Y.; funding acquisition, Z.Z. All authors have read and agreed to the published version of the manuscript.

Funding

This work was financially supported by the Fundamental Research Funds for the Central Universities (WUT: 193108003, 2019IVA032 and No. 215208002) and the Scottish Government’s Rural and Environment Science and Analytical Service Division (RESAS).

Data Availability Statement

Not applicable.

Conflicts of Interest

The authors declare no conflict of interest.

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Catalysts 13 00135 g001 550
Figure 1. The output of research publications using the EC-PS degradation system by the end of 2022. Searched terms: “electrochemistry” and “persulfate”. Source: Web of Science (December 2022).
Figure 1. The output of research publications using the EC-PS degradation system by the end of 2022. Searched terms: “electrochemistry” and “persulfate”. Source: Web of Science (December 2022).
Catalysts 13 00135 g001
Catalysts 13 00135 g002 550
Figure 2. Possible mechanism for the removal of organic contaminants using EPOPs.
Figure 2. Possible mechanism for the removal of organic contaminants using EPOPs.
Catalysts 13 00135 g002
Table
Table 1. Synergistic effects of EC and PS on organic contaminant removal.
Table 1. Synergistic effects of EC and PS on organic contaminant removal.
SystemAnalytesAnalyte Conc.Removal Efficiency (%/min)Ref.
PS/Contact TimeEC/Contact TimeEC + PS/Contact Time
Fe-C/PS2,4-dinitrotoluene100 mg/L≈20/340-94/340[41]
EC/Fe2+/PDSAcid Orange 70.1 mM0/6065.8/6071.6/60[23]
EC/Catalyst/PDSAcid Orange 70.14 mM67.4/30-67.4/30[42]
EC/PDSAniline60 mg/L18/42021/42055/420[43]
EC/US/PDSAniline75 mg/L22/420-69/420[44]
EC/PDSAniline0.45 mM20/150-98/150[45]
EC/PDSAtrazine5 μM0/3036.3/3078.2/30[31]
EC/PDSAtrazine2 μM<10/20-90.6/10[46]
EC/Fe3+/PDSBisphenol A0.22 mM0/600/6026.3/60[29]
EC/Catalyst/PDSBisphenol A0.1 mM4.1/601.8/607.6/60[47]
EC/Catalyst/PMSBisphenol A50 μM0/808/8041.6/80[48]
EC/PDSCarbamazepine5 μM9/3082/3088/20[49]
EC/PDSCarbamazepine5 μM-65/6081/30[50]
EC/Fe3+/PMSCarbamazepine0.04mM1.9/302.18/3012.68/30[39]
EC/PDSCarbamazepine0.042 mM6.87/3043.04/3098.78/30[51]
EC/PDSCiprofloxacin30 mg/L0/80≈40/8096/80[52]
EC/PDSCiprofloxacin20 mg/L0/16025/16090.37/160[53]
EC/PDSDiuron10 μM0/1515/1577/15[40]
EC/PDSDinitrotoluenes300 mg/L
(TOC)		4/48045/48070/480[54]
EC/Catalyst/PDSOrange Ⅱ100 mg/L0/6089.4/6091.2/60[28]
EC/PDSOxcarbazepine20 μM<1/109/1075/10[30]
EC/FeCl2/PDSPhenol100 mg/L0/60-<5/60[55]
EC/PDSSulfamethoxazole5 μM-58/3096/30[56]
EC/PDSTetracycline hydrochloride50 mg/L19.8/24043.8/240≈81/240[57]
Table
Table 2. Effects of current density on degradation efficiencies of organic contaminants in EPOPs.
Table 2. Effects of current density on degradation efficiencies of organic contaminants in EPOPs.
AnalytesSelected Current DensitiesOptimal Current for Higher Removal RatesRef.
1-Butyl-1-methylpyrrolidinium chloride25, 50, 100, 150 mA150 mA[77]
2,4,6-Trichlorophenol50, 60, 70, 80, 90 mA90 mA[78]
2, 4-Dinitrophenol2, 4, 6, 8, 10 mA·cm−24 mA·cm−2[79]
Acid Orange 78.4, 16.8, 33.6 mA·cm−233.6 mA·cm−2[23]
Acid Orange 72, 4, 12, 16 mA·cm−216 mA·cm−2[42]
Ampicillin5, 10, 25, 110 mA·cm−2110 mA·cm−2[70]
Atrazine1, 3, 5, 7, 9 mA·cm−29 mA·cm−2[31]
Atrazine4, 8, 12, 16, 20, 24 mA·cm−224 mA·cm−2[80]
Basic Red 18100, 200, 300, 400 mA400 mA[81]
Bisphenol A1.68, 3.36, 5.04, 8.40 mA·cm−28.40 mA·cm−2[47]
Bisphenol A8.4, 16.8, 33.6 mA·cm−233.6 mA·cm−2[29]
Bisphenol A50, 100, 150 mA·m−2150 mA·m−2[48]
Carbamazepine0.1 × 105, 0.5 × 105, 1 × 105, 2 × 105 mA·m−22 × 105 mA·m−2[50]
Carbamazepine3.57, 7.14, 10.71, 14.28 mA·cm−214.28 mA·cm−2[39]
Ciprofloxacin0.75, 1.45, 2.3 mA·cm−21.45 mA·cm−2[52]
Ciprofloxacin0.9, 1.55, 2.75, 4.25 mA·cm−22.75 mA·cm−2[53]
Cu-EDTA0.1, 0.2, 0.5, 1.0 mA·cm−21.0 mA·cm−2[82]
Diatrizoate1 × 105, 1.5 × 105, 2 × 105 mA·m−22 × 105 mA·m−2[74]
Disperse Blue 35, 10, 20, 30, 40, 80 mA·cm−280 mA·cm−2[83]
Diuron10, 20, 30, 40 mA40 mA[40]
Hexachlorocyclohexanes0, 10, 20 mA20 mA[84]
Orange Ⅱ5.0, 8.4, 11.8, 16.8 mA·cm−216.8 mA·cm−2[85]
Orange Ⅱ1.68, 5.04, 8.40, 11.76, 16.80 mA·cm−216.8 mA·cm−2[28]
Oxcarbazepine1.6–50.0 mA·cm−28.3 mA·cm−2[30]
Phenol1, 3, 5, 10 mA·cm−210 mA·cm−2[61]
Phenol140, 270, 540 mA·L−1540 mA·L−1[60]
Phenol20, 40, 80, 120 mA·cm−2120 mA·cm−2[86]
Pentachlorophenol15, 30, 45, 60, 75, 90 mA90 mA[87]
Reactive Brilliant Blue5, 7.5, 10, 20 mA·cm−220 mA·cm−2[88]
Sulfamethoxazole10, 50, 100, 200 A·m−2200 A·m−2[56]
Sulfamethoxazole5, 10, 15, 20 mA20 mA[89]
Tetracycline hydrochloride6.67, 10, 13.33, 16.67 mA·cm−213.33 mA·cm−2[57]
Trichloroethylene−50, −25, −10, 0, 25, 50, 100 mA100 mA[90]
Table
Table 3. Investigations of optimal pH values in EPOPs.
Table 3. Investigations of optimal pH values in EPOPs.
AnalytesSelected pH ValuesOptimal Ph For Higher Removal RatesRef.
2, 4-Dinitrophenol3, 4, 5, 7, 8, 9, 113[79]
2,4,6-Trichlorophenol3, 5, 7, 9, 119[78]
2,4-dinitrotoluene2, 3, 5, 7, 9, 112[41]
Acid Orange 73, 7, 93[23]
Acid Orange 73, 6, 9, 113[42]
Aniline3, 4, 5, 73[43]
Aniline3, 4, 5, 73[44]
Atrazine4, 6, 8, 104[31]
Atrazine3, 5, 6.3, 7, 9, 113[80]
Bisphenol A3, 6, 93[29]
Bisphenol A3, 6.6, 93[48]
Carbamazepine5, 7, 95[51]
Carbamazepine3, 5, 7, 9, 113[39]
Ciprofloxacin3, 5, 7, 95[52]
Ciprofloxacin3, 5, 7, 97[53]
Dinitrotoluenes0.5, 1, 2, 30.5[54]
Disperse Blue 33, 6.3, 9, 1212[83]
Diuron3, 5, 7, 9, 113[40]
Landfill leachate3.5, 6.5, 8.5, 123.5[93]
Methyl orange3, 5, 7, 93[69]
Orange Ⅱ2, 3, 6, 93[28]
Orange Ⅱ3, 6, 7, 93[85]
Oxcarbazepine3, 5, 7, 9, 113[30]
Pentachlorophenol4.5, 6.5, 8.54.5[87]
Petrochemical wastewater3, 5, 7, 93[94]
Ponceau 6R0.5, 1, 2, 32[95]
Reactive Brilliant Blue3, 6, 9, 116[88]
Sodium dodecylbenzene sulfonate3, 6, 7, 9, 113[68]
Sulfamethoxazole1, 2, 3, 43[89]
Tetracycline hydrochloride2, 4, 7, 94[57]
Table
Table 4. Effects of PS concentrations on the degradation efficiencies of organic contaminants in EPOPs.
Table 4. Effects of PS concentrations on the degradation efficiencies of organic contaminants in EPOPs.
AnalytesSelected PS ConcentrationsOptimal PS ConcentrationRef.
2, 4-dinitrophenol1, 2, 4, 5, 6, 8 mM5 mM[79]
2,4-dinitrotoluene0.037, 0.185, 0.37, 1.85, 3.7 mM0.185 mM[41]
Acid Orange 72, 4, 8, 12 mM12 mM[23]
Acid Orange 72.1, 4.2, 8.4 mM8.4 mM[42]
Ampicillin0.42, 1.05, 2.1 mM2.1 mM[70]
Aniline42, 84, 126, 168 mM126 mM[43]
Aniline42, 84, 105, 126 mM105 mM[44]
Aniline1.85, 3.7, 5.55, 7.4 mM7.4 mM[45]
Atrazine0.25, 0.5, 1, 2, 5 mM5 mM[31]
Atrazine0.3, 0.5, 1, 3, 5 mM5 mM[46]
Atrazine1, 2, 3, 4, 4.5 mM4.5 mM[80]
Bisphenol A1, 5, 10, 20 mM10 mM[29]
Bisphenol A5, 10, 15 mM15 mM[48]
Carbamazepine1, 2, 5 mM5 mM[50]
Carbamazepine0.5, 1, 5, 10, 50, 100, 150 mM100 mM[51]
Carbamazepine0.5, 1, 2, 4, 8 mM2 mM[39]
Dinitrotoluenes25.9, 37, 48.1, 63 mM63 mM[54]
Dinitrotoluenes84, 105, 126, 147 mM126 mM[96]
Diuron0.1, 0.25, 0.5, 0.75, 1 mM1 mM[40]
Greywater3, 4.5, 6, 7.5, 9 mM9 mM[97]
Hexachlorocyclohexanes0.5, 1, 2, 4 mM2 mM[84]
Methyl orange4, 6, 8, 10, 20 mM10 mM[69]
Orange Ⅱ4.2, 6.3, 8.4, 12.6, 16.8 mM16.8 mM[28]
Orange Ⅱ4.2, 8.4, 16.8 mM8.4 mM[85]
Orange G0.05, 0.5, 1, 5 mM5 mM[66]
Oxcarbazepine0.1, 0.25, 0.5, 0.75, 1 mM1 mM[30]
Pentachlorophenol33, 49, 69, 82, 99, 115 μM115 μM[87]
Ponceau 6R1, 1.5, 2, 3, 4 mM4 mM[95]
Pulp and paper wastewater2, 4, 6, 8 mM6 mM[98]
Reactive Brilliant Blue1, 4, 5, 10 mM10 mM[88]
Sodium dodecylbenzene sulfonate10, 25, 50, 100 mM25 mM[68]
Sulfamethoxazole0.1, 1, 2, 5 mM5 mM[56]
Tetracycline hydrochloride4.2, 8.4, 12.6, 16.8 mM12.6 mM[57]
Toluene10, 20, 30 mM30 mM[59]
Table
Table 5. The predominant radical reactions in EPOPs.
Table 5. The predominant radical reactions in EPOPs.
ReactionsRef.
Anode:
M + H2O → M(•OH) + H+ + e−[74]
2SO42− → S2O82− + 2e−[25]
SO42− →SO4•− + e−[71]
Cathode:
S2O82− + e−→SO4•− + SO42−[82]
O2 + 2H+ + 2e−→H2O2[51]
H2O2→2•OH[103]
Radicals:
SO4•− + H2O →H+ + SO42− + •OH[51]
SO4•− + OH−→SO42− + •OH[51]
•OH + •OH→H2O2[103]
SO4•− + SO4•−→S2O82−[51]
Note: M means the organic molecule.
Table
Table 6. The energy consumption of EPOPs in previous studies.
Table 6. The energy consumption of EPOPs in previous studies.
SystemAnalytesC initialElectricity Energy Consumption
(kWh/m3)		Ref.
EC/Pyrite/PMS1-Butyl-1-methylpyrrolidinium chloride323.43 mg/L5.45 kWh/m3[77]
EC/AIS/PDS2, 4-Dinitrophenol200 mg/L0.0336 kWh/m3[79]
EC/Nano-Fe@NdFeB/AC/PDS2,4,6-Trichlorophenol10 mg/L7.44 kWh/m3 (TOC)[78]
EC/MnO2/PDSAcid Orange 749 mg/L7.84 kWh/m3[42]
EC/Fe3O4/PDSAcid Orange 725 mg/L8.64 kWh/m3[102]
EC/CuFe2O4/PDSAtrazine9.92 mg/L0.21 kWh/g[80]
EC/Fe3+/PDSCarbamazepine9.45 mg/L0.0788 kWh/m3[39]
EC/Fe/SBA-15/PDSOrange Ⅱ105.1 mg/L9.87 kWh/m3[85]
EC/PDSSulfamethoxazole25.33 mg/L0.04 kWh/m3[89]
EC/PDSTetracycline hydrochloride50 mg/L11.48 kWh/m3[57]
Note: AIS: Annular iron sheet; Nano-Fe@NdFeB: Iron boron/activated carbon nanocomposite; SBA: 15-Mesoporous silica.
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Sun, J.; Zheng, W.; Hu, G.; Liu, F.; Liu, S.; Yang, L.; Zhang, Z. Electrochemically Assisted Persulfate Oxidation of Organic Pollutants in Aqueous Solution: Influences, Mechanisms and Feasibility. Catalysts 2023, 13, 135. https://doi.org/10.3390/catal13010135

AMA Style

Sun J, Zheng W, Hu G, Liu F, Liu S, Yang L, Zhang Z. Electrochemically Assisted Persulfate Oxidation of Organic Pollutants in Aqueous Solution: Influences, Mechanisms and Feasibility. Catalysts. 2023; 13(1):135. https://doi.org/10.3390/catal13010135

Chicago/Turabian Style

Sun, Jianting, Wei Zheng, Gang Hu, Fan Liu, Siyuan Liu, Lie Yang, and Zulin Zhang. 2023. "Electrochemically Assisted Persulfate Oxidation of Organic Pollutants in Aqueous Solution: Influences, Mechanisms and Feasibility" Catalysts 13, no. 1: 135. https://doi.org/10.3390/catal13010135

APA Style

Sun, J., Zheng, W., Hu, G., Liu, F., Liu, S., Yang, L., & Zhang, Z. (2023). Electrochemically Assisted Persulfate Oxidation of Organic Pollutants in Aqueous Solution: Influences, Mechanisms and Feasibility. Catalysts, 13(1), 135. https://doi.org/10.3390/catal13010135

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