Characteristics and mechanism of electrochemical peroxymonosulfate activation by a Co–N@CF anode for pollutant removal

Abstract

Herein, carbon felt (CF) modified with N and Co (Co–N@CF) was prepared as an anode to effectively activate peroxymonosulfate (PMS) for tetracycline (TC) removal. The influencing factors, kinetic characteristics and the transformation of generated active radicals were further investigated. The results show that N and Co modification of CF via a galvanostatic electrodeposition method results in a significant decrease in the charge transfer resistance (R_ct) of the CF. The fabricated Co–N@CF anode exhibits high catalytic performance and stability towards PMS activation, wherein the TC removal efficiency reached 92.2% within 1 h under optimal conditions. Radical quenching tests and electron paramagnetic resonance spectroscopy experiments demonstrated that SO₄˙⁻, OH˙ and O₂˙⁻ are the main contributors towards TC removal, while ¹O₂ generation is inhibited significantly. The PMS decomposition process driven by the activator enhances the generation of OH˙. The electricity energy demand per order (EE/O) in the reaction system (Co–N@CF anode + PMS) is only 0.072 kWh m⁻³, which makes it an energy-saving approach for pollutant removal. The newly developed fabrication of the Co–N@CF anode for PMS activation could provide insights into the SO₄˙⁻-based electrochemical advanced oxidation process (EAOP) applied in wastewater treatment.

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Water impact

Antibiotic wastewater pollution is a big threat to water environments. Herein, we report an electrochemical peroxymonosulfate (PMS) activation method to enhance tetracycline degradation. The fabricated Co–N@CF anode for PMS activation shows promise for practical application due to its low energy consumption, good reusability and neutral reaction conditions. This work makes an important breakthrough in electrochemical PMS activation for wastewater treatment.

1. Introduction

The increasing quantity of pollutants discharged into the environment has become a complex challenge and has thus attracted increasing research attention.^1–3 Among those pollutants, tetracycline (TC) is one of the typical antibiotics mainly expelled from pharmaceutical factories, the livestock industry and hospitals, which is not readily decomposed in the natural environment.^4–7 Furthermore, as well as damaging the environment, TC can be accumulated in the human body through the food chain and result in severe damage to human health.⁸ Adsorption methods, biodegradation processes, coagulation and chemical processes have been applied to remove antibiotics from the environment.^9–13 However, antibiotics are difficult to remove from the environment using traditional technologies due to their complex chemical structures and poor biodegradability.¹⁴ Advanced oxidation processes (AOPs) are feasible methods by which to degrade antibiotics via the generation of active free radicals.^15–17 Compared to conventional Fenton-like processes, AOP-based sulfate radicals (SO₄˙⁻) generated via persulfate (PS) activation have drawn an increasing amount of focus in pollutant removal considering the long life of the radicals generated in this process, as well as the wide pH range and specific selectivity.¹⁸ To achieve a high yield of sulfate radicals, the application of PS activator with enhanced performance has become an indispensable issue.¹⁹

There are numerous ways to active persulfate to produce sulfate radicals.^13,20–24 Recently, transition-metal activation has received more research attention owing to its energy-saving potential, low cost and high efficiency.²⁵ Transition metals exhibit favorable performance in activating persulfate because they readily give up their electrons to break the O–O bond of PS such as peroxodisulfate (PDS) and peroxymonosulfate (PMS). Generally, PMS with its asymmetric structure can be activated more easily compared to PDS¹³. It has been observed that PMS exhibits a higher catalytic performance than PDS.²⁶ Among the transition metals, cobalt is considered to be the most effective metal to activate PS.^22,27 Therefore, various cobalt-based heterogeneous catalysts have been fabricated that exhibit high efficiency in the removal of organic pollutants via the PS activation process.^13,28–32 However, secondary pollution caused by heavy-metal ion leakage has adverse effects on human health.³³ Thus, an environmentally-friendly and economical activator with high performance should be synthesized. It has been reported that cobalt oxide loaded on activated carbon shows excellent performance in PS activation, exhibiting higher activity than an unsupported heterogeneous Co₃O₄ catalyst and homogeneous Co²⁺ ions.³⁴ However, most methods of activation have restricted implementation due to their high energy demand, high cost and complex equipment management.

Due to its high performance, easy operation and environmental safety, electrochemical PS activation is a promising way to produce radicals and can be applied in AOPs.^35,36 Compared with cathodic activation, anodic activation is a more effective method to use to activate PMS to generate radicals,³⁷ as PMS is negatively charged and there are electrostatic attractions between PMS molecules and the anode surface.³⁸ Carbon-based materials are widely used as substrates for metal nanoparticles due to their stability, good electrical conductivity and low cost.³⁹ The dispersion of cobalt on carbon felt (CF) via chemical binding reduces metal leaching, promotes electron transport and enhances catalytic activity.⁴⁰ Nitrogen with its suitable atomic size and high valence electrons can tightly bind to carbon atoms.⁴¹ Many studies have reported the use of nitrogen-doped carbon materials in electrochemical applications.^42–44 Thus, nitrogen-doped carbon materials could be used as suitable substrates in the electrochemical activation of PMS. Furthermore, reports have shown that single oxygen (¹O₂) is the main active species during the process of cathodic PMS activation,⁴⁵ while in the case of anode oxidation, SO₄˙⁻ is the dominant reactive radical.¹³ Nevertheless, few studies have focused on the degrading of pollutants by cobalt loaded on CF via the assistance of electricity, therefore, the mechanism of PS activation and the transformation of various generated radicals also need further investigation.

Herein, an N- and Co-doped CF (Co–N@CF) was fabricated via a simple electrodeposition and calcination method and applied as an anode for the electrochemical activation of PMS. Dropcasting, hydrothermal and electrodeposition methods are commonly used to fabricate electrodes. The dropcasting method requires a binder to combine catalysts, thus reducing the conductivity of the electrode,⁴⁶ and the hydrothermal method consumes a large quantity of energy.⁴⁷ Thus, electrodeposition was selected to prepare a Co–N@CF anode. The results show that the Co–N@CF anode exhibits good performance towards pollutant removal under optimal conditions. Furthermore, the transformation of the generated radicals during PMS activation was studied via EPR tests and quenching experiments, and subsequently, the main oxidative radicals that contribute towards TC removal were also verified. Our results provide a feasible method to fabricate a CF-based anode to effectively activate PMS and could be applied in wastewater treatment.

2. Experimental section

2.1 Materials

All of the chemical reagents were purchased and used without any further purification. Sodium sulfate (Na₂SO₄), sodium hydroxide (NaOH), cobalt nitrate hexahydrate (Co(NO₃)₂·6H₂O), and sodium thiosulfate (Na₂S₂O₃) were purchased from Tianjin Kemiou Chemical Reagent Co., Ltd. Potassium PMS (KHSO₅), 2-methylimidazole (C₄H₆N₂), p-benzoquinone (C₆H₄O₂) and TC (C₂₂H₂₄N₂O₈) were obtained from Shanghai Macklin Biochemical Co., Ltd. Sulfuric acid (H₂SO₄) and ethanol (CH₃CH₂OH) were purchased from Chengdu Chemical Reagent Co., Ltd. tert-Butyl alcohol ((CH₃)₃COH) and L-histidine (C₆H₉N₃O₂) were provided by Fuchen Chemical Reagent Co., Ltd.

2.2 Preparation of the different anodes by cyclic voltammetry and galvanostatic electrodeposition methods

The CF was cut into several pieces with dimensions of 2 cm × 2 cm. Then, these CF pieces were ultrasonically cleaned in 0.1 M HCl, ethanol and deionized water, sequentially, and dried in an oven at 60 °C overnight. A schematic diagram of the preparation process of the different anodes is shown in Fig. S1.† The preparations of the different anodes by cyclic voltammetry or galvanostatic electrodeposition methods were conducted in a typical three-electrode system with CF, Pt plate and Ag/AgCl as the working, counter and reference electrodes, respectively. A Co@CF anode was prepared via a galvanostatic electrodeposition method with CF poised on a potential of −1.0 V for 1200 s using an electrochemical workstation (CHI-600E, Chenghua, China) in 0.1 M Co(NO₃)₂·6H₂O solution. As a comparison, a Co@CF-cv anode was also fabricated via a cyclic voltammetry electrodeposition method using CF poised on a potential ranging from −1.1 V to 0.1 V at a scan rate of 10 mV s⁻¹ for 10 cycles. After electro-deposition, the Co@CF or Co@CF-cv was taken out and washed with deionized water several times, followed by drying in an oven at 60 °C for 3 h.

To obtain Co- and N-modified CF, the as-prepared Co@CF or Co@CF-cv were immersed in 0.25 M 2-methylimidazole ethanol solution for 24 h. Subsequently, the Co@CF and Co@CF-cv were dried in an oven and then put into a tube furnace and annealed at 600 °C for 2 h at a heating rate of 5 °C min⁻¹ under a N₂ atmosphere to obtain Co–N@CF or Co–N@CF-cv, respectively.

2.3 Characterization

The charge transfer resistance (R_ct) values of the fabricated anodes (bare CF, Co@CF, Co@CF-cv, Co–N@CF and Co–N@CF-cv) were determined by electrochemical impedance spectroscopy (EIS) testing and simulated using the ZSimWin software. EIS tests were performed in a typical three-electrode system containing 0.1 M Na₂SO₄ and the frequency was ranged from 0.01 to 100 000 Hz. A scan electron microscope (SEM, JEOL JSM-6510LV) was used to determine the microstructures of the prepared anodes. X-ray electron spectroscopy (XPS) was carried out using Al Kα radiation. The crystal structures were measured using X-ray diffractometry (XRD, Rigaku UltimaIV) with Cu Kα radiation. EPR spectroscopy (Bruker ZMXmicro-6/1) was applied to identify the generation of radical species using 5,5-dimethy1-1-pyrroline (DMPO) and 2,2,6,6-tetramethyl-4-piperidinol (TMP) as the spin-trapping regents. All EPR tests were conducted at room temperature. The experimental parameters of the EPR spectrometer were as follows: (1) center field: 3500.0 G; (2) sweep width 100.0 G; (3) sweep time: 60 s; (4) sample g-factor: 2.00; (5) receiver gain: 30 dB; (6) modulation amplitude: 1.0 G.

2.4 Degradation experiments

All of the degradation experiments were conducted in a 150 mL reactor containing 0.1 mM of TC. The distance between the anode and counter electrode (Pt plate) was 2.0 cm face-to-face in the reactor, with 50 mM Na₂SO₄ as the electrolyte. At selected reaction intervals, the sample was taken out and the TC concentration was determined using a UV-vis spectrophotometer (Shimadzu, UV-1780) at an absorbance wavelength of 357 nm. The concentration of TC was determined according to a concentration–absorbance standard curve (Fig. S2†). The reaction rate constant (k) was calculated according to the following equation:where C₀, t, C_t and k are initial concentration of TC, the reaction time, the concentration at the selected reaction time and the reaction rate constant, respectively.

3. Results and discussion

3.1 Characterization of the Co- and N-doped CF

Fig. 1b shows the charge transfer resistance (R_ct) values of bare CF, Co@CF, Co@CF-cv, Co–N@CF and Co–N@CF-cv obtained from the EIS tests shown in Fig. 1a. The results show that the R_ct of Co–N@CF is the lowest among the different fabricated anodes and the Co modification of CF significantly reduces the R_ct value of bare CF (Fig. 1b). The R_ct value of Co@CF is lower than that of Co@CF-cv, which indicates that the galvanostatic electrodeposition method is more effective for the electrodeposition of Co on CF than the cyclic voltammetry electrodeposition method. Owing to the doping of N, the electronic arrangement of the original carbon structure is improved, which results in N-doped anodes exhibiting higher electrochemical performance than non-N-doped materials.^47,48 Thus, Co- and N-doping could further significantly reduce the R_ct value of bare CF (Fig. 1b), resulting in the high electrochemical activity of the Co–N@CF anode. Therefore, Co–N@CF was chosen as the anode for PMS activation in the following experiments.

Fig. 1 (a) EIS tests and fitting curves and (b) the R_ct values of different fabricated anodes.

The surface of bare CF can be observed to be clean and smooth (Fig. 2(a–c)), while the Co–N@CF is covered by a thin film (Fig. 2(d–f)), indicating that Co and N were deposited on the surface of the CF. Furthermore, the morphology of the Co–N@CF remained almost unchanged after four consecutive usages of the anode (Fig. 2(g–i)), indicating its favorable stability.

Fig. 2 SEM images of (a–c) bare CF, (d–f) Co–N@CF and (g–i) Co–N@CF after four consecutive uses.

The high-resolution XPS C 1s and O 1s spectra of bare CF are shown in Fig. S3,† where it can be seen that the C 1s spectrum can be well fitted to three peaks at 284.80 eV, 285.8 eV, 288.33 eV, corresponding to C–C, C–O and O–C=O, respectively (Fig. S3a†) and in the high-resolution O 1s spectrum peaks are located at 531.19 eV and 532.60 eV, which correspond to C–O and C=O (Fig. S3b†). Fig. 3a shows the XPS survey spectrum of the Co- and N-doped material compared with that of the bare CF (Fig. 3a), demonstrating that Co and N were doped on the CF surface successfully, and the elemental contents of Co and N were 7.33% and 15.76% in Co–N@CF, respectively (Table S1†). The coexistence of Co and N makes the activation of PMS more efficient via the promotion of electron transfer.⁴⁹Fig. 3b shows the Co 2p high-resolution spectrum, where the main peaks of Co 2p are located at 780.90 and 796.10 eV, corresponding to Co 2p_3/2 and Co 2p_1/2.⁵⁰ The fitting peaks of Co 2p_3/2 were deconvoluted at 778.31, 780.12 and 783.17 eV, which can be attributed to Co, Co³⁺ and Co²⁺, respectively. In addition, there is a satellite peak located at 786.03 eV in high-resolution Co 2p spectrum, which can be ascribed to the shakeup excitation of high-spin Co²⁺. The high-resolution C 1s spectrum could be well fitted well into three different peaks at 284.80, 285.64, and 287.85 eV, corresponding to C–C, C–O–C and O–C=O, respectively (Fig. 3c). Fig. 3d shows the high-resolution N 1s spectrum, which is split into four peaks of nitrogen at 398.40, 399.07, 401.10 and 404.81 eV, corresponding to pyridinic N, pyrrolic N, graphite N and oxidized N, respectively.⁵¹ Generally, pyridinic N (accounting for 7.94% of nitrogen) and pyrrolic N (accounting for 25.37% of nitrogen) with lone-pair electrons for metal coordination combine with cobalt.⁵² In the high-resolution spectrum of O 1s, the three peaks located at 529.87, 531.04 and 532.98 eV correspond to Co–O, C–O, C=O, respectively (Fig. 3e). The XRD analysis shows a clear diffraction peak at 26.4° in the diffraction patterns of CF and Co–N@CF, which is a typical peak of graphitic carbon.⁵³ The XRD peaks of Co–N@CF located at 26.4°, 36.5°,42.4°,61.5°,73.6° and 77.5° can be ascribed to the (002) planes of carbon, and the (111), (200), (220), (311) and (222) planes of CoO, respectively (Fig. 3f). These results indicate that CoO was electrodeposited on N-doped CF successfully, consistent with the XPS analysis results.

Fig. 3 XPS spectra of bare CF and Co–N@CF. (a) Survey spectra of bare CF and Co–N@CF; (b)–(e) XPS spectra of Co–N@CF, for (b) Co 2p, (c) C 1s, (d) N 1s, and (e) O 1s; (f) XRD patterns of bare CF and Co–N@CF.

3.2 Anode performance in the TC degradation process

As shown in Fig. 4a, the degradation efficiency of TC reached 55.6% within 1 h owing to the self-decomposition of PMS and generated strong oxidized radicals to decompose TC. In the presence of CF, the TC removal efficiency increased to 62.3% which was attributed to the additional adsorption of CF. While, in the presence of PMS activator of Co–N@CF, the TC removal efficiency was elevated to 72.2%, indicating the enhanced ability toward PMS activation of CF modified by N and Co.

Fig. 4 (a) Degradation efficiency of TC and (b) the corresponding reaction rate constants of the different reaction systems.

To further investigate the characteristics of the electrochemical activation of PMS, the CF and Co–N@CF anodes were introduced. Using the Co–N@CF anode as the PMS activator (EC/Co–N@CF), the TC removal efficiency was enhanced significantly to 92.2% and the reaction kinetics rate constant (k) was 4.5 times higher than that of the CF anode (EC/CF) (Fig. 4b), which showed that the modification of CF by N and Co contributed towards improving the activation performance of the CF anode. In the EC/Co–N@CF process, radicals such as SO₄˙⁻ and active oxygen radicals (OH˙ and ¹O₂) were generated (Fig. S4†).

3.3 Influencing factors of the degradation of TC and the recyclability of the Co–N@CF anode

As shown in Fig. 5a, the TC degradation efficiencies increased with an increase in the content of PMS and were further enhanced when PMS was activated by the Co–N@CF anode. For example, the degradation efficiency of TC reached 92.2% in a reaction system containing 0.4 mM of PMS activated by the Co–N@CF anode (Fig. 5b), which was 36.6% higher than that in the absence of the Co–N@CF anode. Excess PMS resulted in poor TC removal efficiency, which was also noted (Fig. 5b). This was attributed to the too high amount of PMS consuming SO₄˙⁻ to generate lower activity radicals (SO₅˙⁻).⁵⁴ The reaction stoichiometric efficiency (RSE) of EC/Co–N@CF was 32% according to eqn (2):where Δn(TC) and Δn(PMS) represent the change in the number of moles of TC and PMS, respectively.^55,56

Fig. 5 Degradation efficiency of TC in a reaction system containing different concentrations of PMS in the (a) absence or (b) presence of the Co–N@CF anode and the degradation efficiency of TC influenced by (c) different anode current densities and (d) pH.

The maximum degradation efficiency of TC was observed when the anode current density was 1 mA cm⁻², but a further increase in the anode current density resulted in a decrease in the degradation efficiency (Fig. 5c). There are two possible reasons that might explain this phenomenon, one being that a high current density may cause hydrolysis, and the generated H₂ bubbles attach to the surface of the Co–N@CF anode, inhibiting the reactant mass transfer between the anode surface and bulk solution. The other is that it is likely that the produced radicals such as SO₄˙⁻ or OH˙ are quenched by the anode upon an increase in the applied current density.⁵⁷

Aside from the PMS concentration and current density, pH also plays an important role in TC removal. The results showed that the appropriate pH value for the TC degradation was 7.0, implying that there was no need to add additional reagents to adjust the pH (Fig. 5d), which also indicates the practical application prospects of this system for use in real wastewater treatment. Under acidic conditions at low pH, excess H⁺ consumes SO₄˙⁻ and OH˙ according to eqn (3) and (4).⁵⁸ While, under alkaline conditions, TC is negatively charged, resulting in electrostatic repulsion and worse adsorption toward radicals.⁴⁶ The influence of the initial concentration of TC on the pollutant removal efficiency was also investigated, and the results showed that too high a TC concentration may generate more intermediates to consume free radicals, resulting in poor removal efficiency (Fig. S5†).

OH˙ + H⁺ + e⁻ → H₂O (3)

SO₄˙⁻ + H⁺ + e⁻ → HSO₄˙⁻ (4)

The TC removal efficiency remained almost constant after four consecutive usages (Fig. S6a†) and the kinetic reaction constant also remained at around 0.05 min⁻¹ (Fig. S6b†), indicating that the Co–N@CF anode exhibits desirable stability and recyclability. The electricity energy demand per order (EE/O) was calculated using eqn (5) (an illustration of this equation is shown in Text S1†). The EE/O value in the reaction system containing PMS activated by the Co–N@CF anode was only 0.072 kWh m⁻³ (Table S2†), which was 9.56 and 3.36 times lower than those of the reaction system equipped with the Co–N@CF anode in the absence of PMS and reaction system containing PMS activated by the CF anode, respectively. While, in other research, such as that on an EC/ACF/PMS process, the EE/O was 1.89 kWh m⁻³,⁴⁵ which was 25.25 times higher than the value in our current work (EC/Co–N@CF/PMS). This result shows that the Co–N@CF anode is an energy-saving catalyst, which is beneficial for its practical application.

3.4 Kinetic characteristics of the calculation of the activation energy

Compared with those in the reaction systems containing PMS activated by Co–N@CF, the TC removal efficiency and reaction rate constant (k) were significantly improved at different temperatures (20, 30 and 40 °C) in the reaction systems containing PMS activated by the Co–N@CF anode (Fig. S7† and 6a and b), indicating that PMS can be effectively activated by the Co–N@CF anode and that consequently the pollutant removal efficiency was enhanced.

Fig. 6 Linear fitted curve of the degradation results at various temperatures for the reaction systems containing PMS activated by the (a) Co–N@CF anode and (b) Co–N@CF, (c) the relationship between ln k and 1/T, and (d) the calculated energy demanded during PMS activation by the Co–N@CF anode and Co–N@CF.

The activation energies calculated using eqn (6) (detailed information about eqn (6) is shown in Text S2†) for the reaction systems containing PMS activated by the Co–N@CF anode and Co–N@CF were 37.23 kJ mol⁻¹ and 57.53 kJ mol⁻¹ (Fig. 6d), respectively. This indicates the energy demanded during the PMS activation process by the Co–N@CF anode was 20.3 kJ mol⁻¹ lower than that compared with in the presence of Co–N@CF, resulting in an increased reaction rate. The activation energy of the Co–N@CF anode is much lower than those of other cobalt-based carbon-support catalysts.^59,60 With the synergistic effect of Co and N modification and being electricity driven, the PMS activation becomes more kinetically favorable.

3.5 The proposed mechanism of TC degradation

To verify the main reactive radicals generated during the process of PMS activation by the Co–N@CF anode, TBA, p-BQ and L-histidine were selected as quenchers for OH˙, O₂˙⁻ and ¹O₂,^47,51 respectively, while Na₂S₂O₃ was chosen as a scavenger for the radicals OH˙ and SO₄˙⁻. The reaction rate constants in the presence of the radical quenchers are shown in Table S3.†Fig. 7a shows that the TC removal efficiency was inhibited and decreased to 40.3% in the presence of 100 mM of TBA, and that k was also significantly decreased to 0.01533 min⁻¹ (Table S4†), which indicated the existence of OH˙. Since L-histidine quenches both OH˙ and ¹O₂,⁶¹ the results showed a decreased removal efficiency of TC with increasing L-histidine concentration (Fig. 7b), indicating that both OH˙ and ¹O₂ were the main generated radicals. As shown in Fig. 7c, the TC removal efficiency decreased significantly from 92.2% to 33.5% in the reaction system spiked with 1 mM of p-BQ (O₂˙⁻ quencher), suggesting that O₂˙⁻ may play an important role in TC degradation. The degradation efficiency of TC was inhibited significantly and decreased to only 15.25% in the reaction system spiked with 1 mM of Na₂S₂O₃ (Fig. 7d), which was caused by the presence of Na₂S₂O₃, a scavenger for OH˙ and SO₄˙⁻, showing that these are the active radicals (Fig. 7d). Based on the aforementioned information, radicals including OH˙ and SO₄˙⁻ as well as O₂˙⁻ were generated in PMS activation and contributed towards the degradation of TC.

Fig. 7 (a) Effects of TBA, (b) L-histidine, (c) p-BQ and (d) Na₂S₂O₃ as the different quenching reagents on the pollutant removal efficiency.

EPR tests were further conducted to investigate the generated radicals and their transformation under different conditions. In Fig. 8a, a characteristic 4-line spectrum with an intensity ratio of 1 : 2 : 2 : 1 can be observed, which is typical of a DMPO–OH˙ adduct (hyperfine splitting constants of α_H = α_N = 14.7 G),⁶² indicating the presence of OH˙ in the reaction system.⁶³ Besides this, a typical DMPO–SO₄˙⁻ (hyperfine splitting constants of α_N = 13.2 G, α_H = 9.6, 1.48, 0.78 G) signal was also detected,⁶⁴ which is adjacent to the EPR peaks for OH˙. Meantime, a signal for TMP–¹O₂ with a typical 3-line spectrum in an intensity ratio of 1 : 1 : 1 can also be observed in Fig. 8b. This demonstrates that OH˙, SO₄˙⁻ and ¹O₂ were generated via the PMS activation process. While, the DMPO–O₂˙⁻ adduct was not observed in the EPR tests, which could be ascribed to its kinetic instability and thus ready transformation to a DMPO–OH˙ adduct.⁶⁵ Furthermore, the time course of the EPR tests showed that the peak intensities of the generated radicals (OH˙, SO₄˙⁻ and ¹O₂) decreased with increasing reaction time (Fig. 8).

Fig. 8 EPR spectra of DMPO–OH˙ (), DMPO–SO₄˙⁻ () (a) and TMP–¹O₂ (♠) (b) obtained from the reaction system containing PMS activated by the Co–N@CF anode at different reaction times. Reaction conditions: T = 298 K; PMS = 0.4 mM; current density = 1 mA cm⁻²; pH = 7; Na₂SO₄ = 50 mM.

As previously reported, PMS can self-decompose to produce SO₄˙⁻, OH˙, O₂˙⁻ and ¹O₂ according to eqn (7)–(10).⁶⁶ As shown in Fig. 9a and g, when only PMS was spiked in the reaction system, signals for DMPO–SO₄˙⁻, DMPO–OH˙ and TMP–¹O₂ were observed in the EPR spectrum, confirming that the radicals (SO₄˙⁻, OH˙, ¹O₂) were generated during the PMS self-decomposition process. While, in the presence of Co–N@CF, the EPR peak intensity of DMPO–OH˙ (Fig. 9b) and TMP–¹O₂ (Fig. 9h) were enhanced, indicating that Co–N@CF efficiently activates PMS (eqn (11)). When PMS was activated by the Co–N@CF anode, the EPR signal intensity of DMPO–OH˙ and DMPO–SO₄˙⁻ remained almost constant (Fig. 9c), while the signal intensity of the TMP–¹O₂ adduct decreased significantly (Fig. 9i) compared with that activated by Co–N@CF (Fig. 9b and h). Thus, it was proposed that ¹O₂ generation was inhibited during the process of PMS activation by the Co–N@CF anode (Fig. 10). However, the degradation efficiency and rate were significantly improved in the reaction system containing the Co–N@CF anode, thus, we speculated that SO₄˙⁻ and OH˙ but not ¹O₂ were the predominant radicals that contribute towards pollutant degradation.

Fig. 9 EPR spectra of DMPO–OH˙ (), DMPO–SO₄˙⁻ () (a–f) and TMP–¹O₂ (♠) (g–l) obtained from the reaction system containing PMS with self-decomposition or activated by the Co–N@CF anode and Co–N@CF under different reaction conditions.

Fig. 10 Schematic diagram of PMS activation by the Co–N@CF anode for pollutant removal.

To determine the transformation pathways of O₂˙⁻ and OH˙, the following EPR experiments were conducted. In the reaction system of EC/Co–N@CF with the addition of 1 mM of BQ (O₂˙⁻ quencher), the EPR peak intensity of the DMPO–OH˙ adduct was decreased (Fig. 9e), while the TMP–¹O₂ peak intensity was almost unchanged (Fig. 9k) compared with those in the absence of BQ (Fig. 9c and i). These results illustrate that the transformation of O₂˙⁻ to OH˙ (eqn (12)) was inhibited because of the shortage of O₂˙⁻ in the presence of BQ (O₂˙⁻ quencher) and that the presence of O₂˙⁻ makes no contribution towards ¹O₂ generation (eqn (13)). Furthermore, the self- or activator-driven decomposition of PMS could be considered as the main pathways to produce OH˙ (eqn (7) and (14)), which was further proved by the EPR tests in reaction systems spiked with MeOH, and the results showed that the EPR peak of the DMPO–OH˙ adduct was almost not detectable (Fig. 9f). This is as a result of the shortage of SO₄˙⁻ in the presence of MeOH (the SO₄˙⁻ quencher),⁶⁷ confirming that OH˙ should be derived from the hydrolysis of SO₄˙⁻ in aqueous solution (eqn (14)). Besides this, a negligible amount of OH˙ and H₂O₂ may be generated on the surfaces of the Co–N@CF anode and counter electrode according to eqn (15) and (16), respectively.^45,47 As previous reported, O₂˙⁻ is mainly generated by PMS self-decomposition (eqn (8)–(10)) or O₂ reduction (eqn (17)).⁶⁸

Meanwhile, in the reaction system of EC/Co–N@CF saturated by N₂, the EPR peak intensity of the DMPO–OH˙ (Fig. 9d) and TMP–¹O₂ adduct (Fig. 9j) remained almost unchanged, considering the absence of O₂ in the reaction system saturated by N₂, with the results showing that the transformation of the oxygen-containing radicals through the pathways for O₂ reduction to generate O₂˙⁻ and ¹O₂ (eqn (17) and (18)) could be neglectable (Fig. 10). Therefore, we can conclude that O₂˙⁻ was mainly generated by PMS self-decomposition.

HSO₅⁻ + SO₅²⁻ → HSO₄⁻ + SO₄˙⁻ + ¹O₂ (7)

HSO₅⁻ + H₂O → HSO₄⁻ + H₂O₂ (8)

H₂O₂ + OH˙ → HO₂˙ + H₂O (9)

HO₂˙ → H⁺ + O₂˙⁻ (10)

O₂˙⁻ + H₂O₂ → OH˙ + OH⁻ + O₂ (12)

O₂˙⁻ + OH˙ → ¹O₂ (13)

SO₄˙⁻ + H₂O → OH˙ + SO₄²⁻ + H⁺ (14)

O₂ + 2H⁺ + 2e⁻ → H₂O₂ (15)

H₂O → OH˙ + H⁺ + e⁻ (16)

O₂ + e⁻ → O₂˙⁻ (17)

4. Conclusion

Herein, a Co–N@CF anode was fabricated via a simple electrodeposition method and further applied to efficiently activate PMS to degrade TC. Under optimal conditions, the TC degradation efficiency reached 92.2% within 1 h using the Co–N@CF anode as the PMS activator. N and Co modification of the CF via the galvanostatic electrodeposition method significantly decreased the R_ct value of CF, which contributed towards the high PMS activation performance of the fabricated Co–N@CF anode. Quenching experiments and EPR tests demonstrated that OH˙, O₂˙⁻ and SO₄˙⁻ were the main contributors for pollutant degradation. More importantly, we found that the generation of ¹O₂ was significantly inhibited during the PMS activation process by the Co–N@CF anode. The facile prepared Co–N@CF anode with good PMS activation performance exhibits practical application prospects due to low cost and low energy consumption, good reusability and no need for pH adjustment (neutral reaction conditions). Our current study offers a strategy for the electroactivation of PMS, which could be applied in SO₄˙⁻-based advanced oxidation processes.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This work was financially supported by the Strategic Research and Development Program of Shaanxi Province (2019ZDLSF06-02) and the Special Project for Serving Local S&T Development of Education Department, Shaanxi Province (21JC022).

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Footnote

† Electronic supplementary information (ESI) available. See DOI: 10.1039/d1ew00676b

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