Amorphous MnO₂ Supported on CN@SiO₂ for Levofloxacin Degradation via a Non-Radical Pathway by PMS Activation

Abstract

Mn-based catalysts have been extensively studied in advanced oxidation processes based on peroxymonosulfate (PMS) oxidants, demonstrating their significant potential for treating antibiotic-contaminated wastewater. In this study, an amorphous MnO₂-based composite catalyst (MnO₂/CN@SiO₂) was prepared and used to activate PMS for degrading levofloxacin (LEV). The effects of reaction conditions, such as reaction temperature, catalyst dosage, PMS concentration, and solution pH, on LEV degradation were comprehensively investigated. The interference of water components, e.g., ${\mathrm{NO}}_3^{-}$, ${\mathrm{SO}}_4^{2-}$, Cl⁻, ${\mathrm{CO}}_3^{2-}$, and humic acid, on the degradation efficiency of LEV was analyzed, and the stability of the catalysts was explored by cycling experiments. Finally, radical quenching experiments and electron paramagnetic resonance spectroscopy were employed to elucidate the contribution of active species to the degradation reaction process. A non-radical-based pathway for LEV degradation was proposed based on these results.

1. Introduction

Antibiotics are widely used in the pharmaceutical and aquaculture industries due to their excellent bacteriostatic properties. With the increasing demand for antibiotics in recent years, substantial amounts of antibiotic effluent from hospitals, animal husbandry, and aquaculture have been discharged into aquatic environments [1,2]. Levofloxacin (LEV), a third-generation quinolone with broad-spectrum antimicrobial properties, is commonly used in the treatment of urinary, gastrointestinal, and respiratory diseases [3,4]. Typically, LEV cannot be fully absorbed and utilized in humans and animals, leading to unmetabolized drugs entering ecosystems. This phenomenon not only exacerbates water pollution, but also accelerates the development of bacterial resistance, thereby posing a great threat to human health [5,6,7]. Therefore, the treatment of antibiotic-containing wastewater has become an urgent priority. However, the chemical stability of antibiotics makes traditional wastewater treatment processes ineffective in removing them. Advanced oxidation processes (AOPs), which possess superior oxidizing capabilities, hold greater promise for treating antibiotic-containing wastewater [8,9].

In advanced oxidation techniques, free radicals with strong oxidizing properties (e.g., •OH, •${\mathrm{SO}}_4^{-}$, and •${\mathrm{O}}_2^{-}$) can effectively degrade antibiotics and achieve a high degree of mineralization of organic pollutants in wastewater [10,11]. Peroxymonosulfate (PMS)-based advanced oxidation processes (SR-AOPs) are commonly used for the treatment of antibiotic pollutants in water due to the high redox potential and long half-life of ${\mathrm{SO}}_4^{•-}$ [12,13,14]. Compared to H₂O₂, solid PMS is easier to store and transport [15]. In summary, SR-AOPs show great promise in antibiotic degradation. Transition metals, such as copper, iron, cobalt, and manganese, can provide electrons to promote the breaking of S–O bonds and are commonly used to activate PMS in heterogeneous systems [16,17,18,19]. Among these metals, Mn is usually used for PMS activation in the degradation of antibiotic wastewater because of its abundant reserves, low environmental toxicity, rich valence composition, and controllable crystal morphology [20]. The different arrangements of the basic structural unit [MnO₆] allow for the existence of a variety of crystals of MnO₂, including α, β, γ, and δ–MnO₂ [21]. In addition, differences in crystalline structure and manganese ion valence states in various MnO₂ types lead to the generation of distinct reactive oxygen species (ROS), which in turn exhibit different catalytic activities [22,23]. Recently, it has been reported that amorphous MnO₂ can rapidly activate PMS to degrade pollutants via a non-radical pathway [24]. Generally, highly crystalline MnO₂ needs to be synthesized under the conditions of hydrothermal, microwave, or roasting methods, which are energy-intensive and complex [25,26]. In contrast, amorphous MnO₂ shows great advantages in preparation. It can be obtained by simple redox reactions involving ${\mathrm{M}\mathrm{n}\mathrm{O}}_4^{-}$ or Mn²⁺ at ambient temperature and pressure [27]. However, MnO₂ particles tend to agglomerate during the preparation process. This agglomeration seriously hinders the exposure of active sites, thereby diminishing the activation efficiency of PMS. Consequently, enhancing the performance of PMS activation necessitates the introduction of additional components to effectively disperse MnO₂.

Graphitic-phase carbon nitride (g–C₃N₄) is a non-metallic organic polymer with high thermal and chemical stability, commonly used as a carrier in material preparation [28]. The bulk structure of g–C₃N₄ consists of a large number of stacked, flattened layers, which inherently lead to a relatively low specific surface area and limited surface functional groups. Thermal exfoliation has been demonstrated to effectively reduce the stacking degree of these layered structures, thereby significantly increasing the specific surface area [29,30]. The –OH content on the surface of Mn-based catalysts has a significant effect on the activation of PMS [31]. SiO₂, as a hydrophilic material rich in –OH groups, provides binding sites for interfacial interactions between materials and enhances the dispersion stability of composites in aqueous solutions [32,33]. As far as we know, there are relatively few studies on improving the activation efficiency of PMS by compositing SiO₂ with MnO₂. SiO₂ rich in –OH groups can accelerate the binding rate of PMS to the catalyst. Furthermore, the stabilized g–C₃N₄ provides a support matrix for the effective dispersion of MnO₂ particles. Therefore, the composite of MnO₂ with SiO₂ and g–C₃N₄ is expected to solve the problems of easy agglomeration of MnO₂ and slow activation of PMS.

In this study, MnO₂/CN@SiO₂ (MCS) ternary composites were successfully synthesized and used for activating PMS to degrade LEV. The physicochemical properties of the catalysts were comprehensively analyzed by XRD, XPS, and SEM techniques. The catalytic performance of catalyst compositions for LEV degradation by activating PMS was systematically compared. The effects of catalyst dosage, reaction temperature, solution pH, and inorganic anion interference on PMS activation efficiency were thoroughly investigated. Furthermore, the stability of the catalyst was evaluated via cycling experiments. Radical quenching experiments and electron paramagnetic resonance (EPR) spectroscopy were employed to unveil the primary degradation pathways during the reaction process. Based on these results, a non-radical pathway-based mechanism for LEV degradation was proposed. The experimental results showed that the MnO₂/CN@SiO₂ catalyst had the ability to activate PMS efficiently.

2. Results

2.1. Characterizations of MnO₂/CN@SiO₂

Figure 1 shows the XRD patterns of MnO₂/SiO₂, MnO₂/CN, MCS, and MnO₂/g-C₃N₄@SiO₂. The broad diffraction peaks at 20–30° can be attributed to amorphous SiO₂ [34]. The peak observed at 27.6° corresponds to the (002) crystallographic plane of g-C₃N₄ (PDF#87-1526) [35]. In addition, the weak diffraction peaks at 37° and 66° are related to amorphous MnO₂ [36]. The presence of SiO₂, CN, and MnO₂ diffraction peaks confirms the successful preparation of the composite. By comparing the XRD patterns of MCS and MnO₂/g–C₃N₄@SiO₂, a significant difference was identified in the intensity of the main diffraction peak at around 23°. This discrepancy may arise from the controlled growth of SiO₂ and MnO₂ after thermal stripping of g–C₃N₄.

The surface elements and valence states of the MCS catalyst before and after the reaction were analyzed using XPS. The full XPS spectra of both the fresh catalyst and the recycled catalysts confirm the presence of Mn, O, N, C, and Si elements (Figure 2a). In Figure 2b, the two peaks located at approximately 642 eV and 654 eV correspond to the Mn 2p_3/2 and Mn 2p_1/2 orbitals, respectively [37]. The high-resolution Mn 2p spectra can be deconvoluted into four peaks: those at 641.3 and 641.6 eV are attributed to Mn³⁺, while those at 642.5 and 642.7 eV are assigned to Mn⁴⁺ [38]. As presented in Figure 2c, the high-resolution C 1s spectrum of the fresh MCS catalyst shows three peaks located at 284.8 eV, 286.1 eV, and 288.6 eV, corresponding to C–C, C–N, and the N–C=N bonds, respectively [39]. The high-resolution N 1s spectra (Figure 2d) can be deconvoluted into three peaks located at 399.1 eV, 399.8 eV, and 401.7 eV, representing sp² hybridized N atoms (N–C=N), quaternary N atoms ((C)₃–N), and N atoms in the amino group (C–NH_x), respectively [38,40]. The high-resolution O 1s spectra of the fresh MCS catalyst (Figure 2e) can be divided into three distinct peaks at 529.8 eV, 531.3 eV, and 532.3 eV, which are attributed to the lattice oxygen (Mn–O) in the metal oxides, adsorbed hydroxyl oxygen (–OH), and oxygen in chemisorbed water on the surface of the material [41]. As shown in Figure 2f, the characteristic peak at 101.9 eV for the fresh catalyst corresponds to the 2p electron orbital of the Si element [42].

Figure 3 demonstrates the SEM image, elemental mapping images, and energy dispersion spectrum of the MCS composites. As shown in Figure 3a, amorphous MnO₂ particles are distributed on the surface of the CN@SiO₂ substrate. The elemental mapping images (Figure 3b) reveal the distribution of the C, N, O, Si, and Mn elements across the substrate surface. Notably, the dispersion of Si is much denser than those of C and N, indicating the existence of a SiO₂-coated CN structure. Furthermore, the significant overlap of Mn, O, and Si elements indicates the successful preparation of the MCS composite.

2.2. Effect of Catalyst Composition on LEV Degradation

The degradation of LEV by different catalysts in the presence of PMS was investigated. As shown in Figure 4a, MnO₂/CN exhibited superior degradation performance compared to MnO₂/g–C₃N₄, which may be attributed to greater exposure of MnO₂ on CN than g–C₃N₄. Moreover, the binary CN/SiO₂ composite-supported catalyst demonstrated higher catalytic activity than the CN-supported catalyst. Figure 4b shows the effect of different MnO₂ contents on the LEV degradation efficiency. At a MnO₂ content of 50 wt%, the LEV degradation efficiency reached 80.8%, which is notably higher than that at 20 wt% (58.7%), 80 wt% (72.3%), and pure MnO₂ (67.4%).

The catalytic activities of the MnO₂/CN, MnO₂/SiO₂, and MCSs with various SiO₂/CN ratios for LEV degradation are presented in Figure 4c. The degradation efficiencies of MnO₂/CN, MCS-1 (SiO₂/CN = 1:4), MCS-2 (SiO₂/CN = 1:1), MCS-3 (SiO₂/CN = 4:1), and MnO₂/SiO₂ within 40 min are 68.2%, 80.8%, 77.0%, 76.9%, and 76.1%, respectively. As depicted in Figure 4d, the LEV degradation efficiencies with PMS alone, MnO₂ alone, and MCS alone are 1.9%, 7.8%, and 11.1%, respectively, indicating that neither PMS nor the catalysts can effectively degrade LEV. In contrast, MnO₂ can effectively activate PMS, leading to a LEV degradation efficiency of 67.4%. The MCS showed higher catalytic activity than the pure MnO₂ catalyst.

2.3. Effect of Reaction Conditions on LEV Degradation

Figure 5 illustrates the effects of catalyst dosage, PMS concentration, initial LEV concentration, and reaction temperature on LEV degradation. As depicted in Figure 5a, the LEV degradation efficiency increased from 67.2% to 83.7%, with a corresponding increase in k_app from 0.0241 min⁻¹ to 0.1354 min⁻¹, as the catalyst dosage was raised from 5 mg to 20 mg. This result is consistent with the expectation that an increase in catalyst dosage can provide more reactive sites, thus accelerating the PMS activation and improving the LEV degradation. As shown in Figure 5b, as the concentration of PMS increased from 0.5 mM to 1.0 mM, the degradation efficiency of LEV within 40 min increased from 75.0% to 80.1%, and the k_app of the reaction system increased from 0.0329 min⁻¹ to 0.0945 min⁻¹. However, further increasing the PMS concentration resulted in insignificant enhancement in LEV degradation due to the limited number of active sites provided by fixed catalyst dosage. The effect of initial LEV concentration on LEV degradation is illustrated in Figure 5c. As expected, the degradation efficiency of LEV decreased from 84.5% to 80.8% with an increase in LEV concentration from 20 mg/L to 50 mg/L. This reduction can be attributed to the fixed catalyst dosage and PMS concentration, which result in a constant production of oxidative species throughout the reaction system. Consequently, under identical conditions, a higher LEV concentration leads to lower degradation efficiency. As displayed in Figure 5d, the degradation efficiency of LEV increased from 69.3% to 80.1% with an increase in reaction temperature from 10 °C to 40 °C. The elevated temperature favors the thermally driven activation of PMS, thereby promoting LEV degradation [43]. Based on the Arrhenius plot, the reaction activation energy was determined to be 18.176 kJ/mol.

Figure 6 illustrates the effect of the initial solution pH on LEV degradation, PMS concentration, and potential Mn ion leaching. It is evident that pH significantly affects LEV degradation efficiency. An approximately 80.1% degradation efficiency was obtained in the pH range of 3–7. In contrast, only 24.8% and 17.7% LEV degradations were achieved at pH = 9 and pH = 11, respectively. Figure 6c presents the variation in PMS concentration with pH in LEV solutions without a catalyst. The PMS concentration was determined using the iodometric method after 10 min of stirring. PMS was stable under acidic and neutral conditions, but its concentration decreased in alkaline environments due to hydrolysis. This can explain the low LEV degradation efficiency at a high pH. At the same time, the effect of pH on catalyst stability was investigated by dispersing the catalyst in LEV solutions with varying pH levels for 10 min. Subsequently, PMS was added to initiate the reaction after filtering out the catalyst. As shown in Figure 6d, substantial LEV degradation occurred under acidic conditions (pH = 3–5), implying the happening of a homogeneous reaction due to leached Mn ions. According to previous studies, under acidic conditions, the catalyst enhances PMS activation via electrostatic attraction. However, as pH increases, the electrostatic effect diminishes or even transitions to electrostatic repulsion, significantly inhibiting PMS activation by the catalyst [44,45,46]. In conclusion, the pH of the solution not only impacts the stability of PMS and the catalyst but also interferes with catalytic PMS activation, thereby substantially influencing the system’s degradation performance.

Inorganic anions commonly found in natural aquatic environments, including ${\mathrm{NO}}_3^{-}$, ${\mathrm{SO}}_4^{2-}$, Cl⁻, and ${\mathrm{CO}}_3^{2-}$, can significantly influence the degradation of LEV. The effects of the four anions on the LEV degradation in the MCS-PMS system were investigated (Figure 7). The results show that Cl⁻ promotes LEV degradation, whereas the other three anions inhibit the reaction to varying degrees, with ${\mathrm{CO}}_3^{2-}$ being the most potent inhibitor. In solution, Cl⁻ reacts with ${\mathrm{SO}}_4^{•-}$ to produce additional active chlorine species (${\mathrm{C}\mathrm{l}}_2^{-}$ and HOCl), which actively participate in the degradation process [47]. Therefore, increasing Cl⁻ concentration promotes LEV degradation efficiency [48]. Conversely, the inhibitory effects of ${\mathrm{NO}}_3^{-}$, ${\mathrm{SO}}_4^{2-}$, and ${\mathrm{CO}}_3^{2-}$ may arise from their reactions with free radicals and/or competition with $\mathrm{H}{\mathrm{SO}}_5^{-}$ for the active site [49]. Notably, the pronounced inhibitory effect of ${\mathrm{CO}}_3^{2-}$ could also stem from its influence on solution pH. Upon the addition of ${\mathrm{CO}}_3^{2-}$, the resulting alkaline conditions hinder PMS activation, thereby reducing LEV degradation efficiency (Figure 6).

Natural organic matter (NOM), ubiquitously present in natural aquatic environments, can inhibit the catalytic degradation process via reactive oxygen scavenging and competitive adsorption [50]. Humic acid (HA), a major component of NOM, interacts with functional groups on the catalyst surface, thereby hindering the binding of the catalyst to PMS and influencing the degradation reaction.

The effects of different HA concentrations on the degradation of LEV in the MnO₂/CN@SiO₂-PMS system were investigated, with the results presented in Figure 8a. When the HA concentration was less than 10 mg/L, its effect on LEV degradation was negligible. However, upon increasing the HA concentration to 20 mg/L, the degradation efficiency decreased significantly, from 80.8% to 72.2%.

The stability of MnO₂/CN@SiO₂ catalysts during degradation was investigated by cycling experiments, as shown in Figure 8b. Multiple replicate experiments were performed under the same conditions. After four cycles, the LEV degradation efficiency decreased from 83.5% to 74.7%. This reduction was likely attributed to the inevitable leaching of Mn ions during the activation of PMS by MnO₂, which consequently reduced the MnO₂ content in the catalyst, thus diminishing its catalytic performance. Despite this, the cycling experiments confirmed the satisfactory stability of the catalysts, with only an 8.5% decline in activity after four cycles.

2.4. Identification of Active Species

Free radicals such as •${\mathrm{SO}}_4^{-}$, •OH, and •${\mathrm{O}}_2^{-}$ are usually considered the main reactive substances in AOPs [51]. To identify the potential reactive substances in the MnO₂/CN@SiO₂-PMS system and their respective contributions to the degradation process, scavenging experiments were conducted. Tert-butyl alcohol (TBA), methanol (MeOH), p-benzoquinone (PBQ), and L-histidine (L-His) were used as scavengers because they can quench •OH, •${\mathrm{SO}}_4^{-}$/•OH, •${\mathrm{O}}_2^{-}$, and ¹O₂, respectively.

Based on the reaction rates of the quencher and the radicals, different quencher concentrations were chosen to ensure that the same radicals were quenched in the same amount of time [52]. Figure 9 shows the effect of various quenchers on the degradation of LEV. Figure 9a shows the experimental results of tert-butanol quenching of •OH radicals in the MnO₂/CN@SiO₂-PMS system. Upon adding TBA at a concentration of 10 mM, the degradation efficiency of LEV decreased from 80.8% to 79.3%. Further increasing the TBA concentration to 50 mM led to a reduction in the LEV degradation efficiency. The experimental results of MeOH quenching of •${\mathrm{SO}}_4^{-}$ and •OH radicals are shown in Figure 9b. When MeOH was added at a concentration of 10 mM, the LEV degradation rate decreased from 80.8% to 78.7%, and it further dropped to 74.6% at a concentration of 50 mM.

Figure 9c illustrates the experimental results of the PBQ quenching system of •${\mathrm{O}}_2^{-}$ radicals. At PBQ concentrations of 1 mM and 5 mM, the LEV degradation efficiencies were reduced to 79.2% and 77.6%, respectively. In summary, the quenching experiments targeting •${\mathrm{SO}}_4^{-}$, •OH, and •${\mathrm{O}}_2^{-}$ showed that the free radical pathway is not the dominant pathway for LEV degradation. Consequently, it is hypothesized that the non-free radical pathway may play a more significant role in the degradation of LEV. The ¹O₂ quenching experiment is shown in Figure 9d. L-His exerted a tremendous inhibitory effect on the degradation of LEV, with the degradation efficiency decreasing to 50.4% and 22.2% at concentrations of 1 mM and 5 mM, respectively. Previous studies have shown that furfural and L-His can not only quench ¹O₂ but also directly react with PMS, thus inhibiting the reaction [53].

Figure 10a demonstrates the changes in PMS concentration after the addition of different scavengers. TBA, MeOH, PBQ, and L-His were added to the PMS solution with an initial concentration of 1 mM, and the concentration changes of PMS were determined using the iodometric method. The experimental results clearly indicate that the PMS concentration decreased markedly upon adding L-His, confirming its ability to quench PMS.

As shown in Figure 10b, the effect of scavengers on solution pH was investigated. The presence of TBA and PBQ can decrease the solution pH, while the presence of MeOH and L-His can increase it. Slight changes in solution pH do not significantly affect the experimental results. Therefore, the inhibition of the reaction by the scavenger through changing the solution pH can be ignored.

To further investigate the role of ¹O₂ in the degradation process, the production of ¹O₂ during the reaction was examined by EPR spectroscopy. In this study, ¹O₂ was captured using 2,2,6,6-tetramethylpiperidine (TEMP) as a spin-trapping agent, and the results are shown in Figure 11a. The MnO₂/CN@SiO₂-PMS system clearly detected the TEMP-¹O₂ triple peak signals with an approximate intensity ratio of 1:1:1, confirming the presence of ¹O₂ in the reaction [54]. Conversely, no significant ¹O₂ signal was detected in the MnO₂-PMS system, yet LEV degradation still occurred, suggesting that ¹O₂-mediated degradation was not the sole pathway for LEV removal. As depicted in Figure 11b, PMS consumption was monitored at a reaction time of 10 min in the catalyst-free system, the MnO₂-PMS system, and the MCS-PMS system. Without the addition of a catalyst, the PMS concentration remained essentially constant. In contrast, in the MnO₂-PMS and MCS-PMS systems, the PMS concentrations decreased to 91.5% and 77.8% of the initial value, respectively. The PMS consumption in both systems was consistent with the ¹O₂ signal intensities observed in the EPR spectra described above, indicating that the MnO₂/CN@SiO₂ catalyst could activate PMS more efficiently to degrade LEV.

2.5. Reaction Mechanism

Based on the above experimental and characterization results, a reaction mechanism for LEV degradation in the MnO₂/CN@SiO₂-PMS system was proposed. A non-radical pathway was proposed as the primary mechanism for LEV degradation. In the MnO₂/CN@SiO₂-PMS degradation system, MnO₂ serves as an activator for PMS, with Mn³⁺ acting as the main active center. The activation of PMS can be categorized into two pathways, which are elaborated as follows. Firstly, PMS binds to the active site of MnO₂ by substituting –OH groups, subsequently reacting with Mn³⁺ to form the complex Mn³⁺–(O)${\mathrm{O}\mathrm{S}\mathrm{O}}_3^{-}$ (Equation (1)) [55]. Electron transfer occurs between MnO₂ and PMS, enabling the complex to extract one electron from LEV, thereby oxidizing it. Simultaneously, Mn³⁺ loses one electron to form Mn⁴⁺, while $\mathrm{H}{\mathrm{SO}}_5^{-}$ gains two electrons to form ${\mathrm{SO}}_4^{2-}$ (Equation (2)). This pathway constitutes the dominant mechanism for LEV degradation in the system [56]. Secondly, $\mathrm{H}{\mathrm{SO}}_5^{-}$ adsorbed onto the catalyst surface reacts directly with Mn³⁺ to form •OH, Mn⁴⁺, and ${\mathrm{SO}}_4^{2-}$, or reacts with Mn⁴⁺ to form Mn³⁺ and ${\mathrm{SO}}_5^{-}$ (Equations (3)–(6)). During the reaction, the reduction in Mn content leads to a decrease in lattice oxygen content (Figure 2e), resulting in the formation of oxygen vacancies and reactive oxygen species (O*). O* reacts with PMS to form ¹O₂ (Equation (7)). In addition, ¹O₂ is produced during the hydrolysis of ${\mathrm{S}\mathrm{O}}_5^{•-}$ and the autolytic decomposition of PMS (Equations (8) and (9)). Ultimately, the ROS generated in the reaction contribute to the degradation of LEV [57].

$$\equiv {\mathrm{Mn}}^{3+}–\mathrm{OH}+{\mathrm{HSO}}_5^{-}\to \equiv {\mathrm{Mn}}^{3+}–\left(\mathrm{O}\right) {\mathrm{OSO}}_3^{-}+{\mathrm{H}}_2\mathrm{O}$$

(1)

$$\equiv {\mathrm{Mn}}^{3+}–\left(\mathrm{O}\right) {\mathrm{OSO}}_3^{-}+{\mathrm{e}}^{-}\to \equiv {\mathrm{Mn}}^{4+}+{\mathrm{SO}}_4^{2-}$$

(2)

$${\mathrm{Mn}}^{3+}+{\mathrm{HSO}}_5^{-}\to {\mathrm{Mn}}^{4+}+{\mathrm{OH}}^{-}+{ \mathrm{SO}}_4^{•-}$$

(3)

$${\mathrm{SO}}_4^{•-}+{\mathrm{H}}_2\mathrm{O} \to •\mathrm{OH}+{ \mathrm{SO}}_4^{2-}+{\mathrm{H}}^{+}$$

(4)

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

(5)

$${\mathrm{Mn}}^{4+}+{\mathrm{HSO}}_5^{-}\to {\mathrm{Mn}}^{3+}+{\mathrm{H}}^{+}+{\mathrm{SO}}_5^{•-}$$

(6)

$$\mathrm{O}*+ {\mathrm{HSO}}_5^{-}\to {\mathrm{HSO}}_4^{-}+{1\mathrm{O}}_2$$

(7)

$${\mathrm{SO}}_5^{•-}+{\mathrm{H}}_2\mathrm{O} \to {\mathrm{HSO}}_4^{-}+{1\mathrm{O}}_2$$

(8)

$$\mathrm{H}{\mathrm{SO}}_5^{-}+{\mathrm{SO}}_5^{2-}\to \mathrm{H}{\mathrm{SO}}_4^{-}+{\mathrm{SO}}_4^{2-}+{1\mathrm{O}}_2$$

(9)

3. Materials and Methods

3.1. Materials

Potassium permanganate (≥99.5% KMnO₄), sodium hydroxide (≥96.0% NaOH), manganese sulfate monohydrate (≥99.0% MnSO₄·H₂O), ammonium persulphate (≥98.0% (NH₄)₂S₂O₈), sodium chloride (≥99.5% NaCl), sodium sulfate (≥99.0% Na₂SO₄), sodium carbonate (≥99.0% Na₂CO₃), sodium nitrate (≥99.0% NaNO₃), ethanol (≥99.5% C₂H₅OH), methanol (≥99.5% CH₃OH), tert-butyl alcohol (≥99.0% TBA), para-benzoquinone (≥99.0% PBQ), sodium bicarbonate (≥99.5% NaHCO₃), and hydrochloric acid (36–38% HCl) were all purchased from Sinopharm Chemical Reagent Co., Ltd. (Beijing, China). Melamine (≥99.0% C₃H₆N₆), tetraethyl orthosilicate (≥99.0% TEOS), levofloxacin (≥99.0% LEV), potassium monopersulfate triple salt (≥42.0% KHSO₅·0.5KHSO₄·0.5K₂SO₄), 2,2,6,6-tetramethylpiperidine (≥98.0% TEMP), and humic acid (≥90.0% HA) were purchased from Shanghai Aladdin Biochemical Technology Co., Ltd. (Shanghai, China). Potassium iodide (≥99.5% KI), and L-histidine (≥99.0% L-His) were purchased from Shanghai Maclean’s Biochemical Technology Co., Ltd. (Shanghai, China).

3.2. Preparation of MnO₂/CN@SiO₂

The MnO₂/CN@SiO₂ catalyst preparation consists of three steps, which are illustrated in Figure 12.

3.2.1. Synthesis of CN

A total of 5.0 g of melamine was added into a crucible with a lid and subsequently transferred to a muffle furnace for pyrolysis. The sample was heated to 550 °C at a rate of 5 °C/min and held at this temperature for 2 h. After natural cooling, yellow solid carbon nitride (g–C₃N₄) was successfully obtained. Thermally stripped carbon nitride (CN) was obtained by reheating g–C₃N₄ in a muffle furnace at 525 °C for 2 h.

3.2.2. Synthesis of CN@SiO₂

A total of 0.06 g of CN was added into a mixture of 22.5 mL of anhydrous ethanol and 7.5 mL of deionized water. Following ultrasonication for 10 min to disperse the CN, 0.045 mL of TEOS was added dropwise under stirring. After continuously stirring for 30 min, 0.2 mL of 6 wt% NaOH solution was added dropwise. Stirring was maintained for 2.5 h to ensure complete hydrolysis of TEOS, thereby obtaining CN@SiO₂. Finally, the product was centrifuged, washed three times with deionized water, and dispersed in 5 mL of deionized water.

3.2.3. Synthesis of MnO₂/CN@SiO₂

A total of 0.1125 g of KMnO₄ was dissolved in 30 mL of deionized water. Subsequently, the CN@SiO₂ was added to the KMnO₄ solution and stirred for 30 min. Then, 4.5 mL of anhydrous ethanol was added dropwise. After continuous stirring for 4 h, the resulting product was centrifuged, washed with deionized water and anhydrous ethanol, and dried at 70 °C to obtain the MnO₂/CN@SiO₂ catalyst.

Meanwhile, MnO₂/g–C₃N₄ and MnO₂/g–C₃N₄@SiO₂ composites were synthesized using g–C₃N₄ according to the above method, distinguishing between MnO₂/CN and MnO₂/CN@SiO₂ prepared with CN.

3.3. Characterizations

X-ray diffraction (XRD) was performed on a D/max-2500/PC model X-ray powder diffractometer (Rigaku, Akishima, Japan) to determine the crystal structure of the catalyst. X-ray photoelectron spectroscopy (XPS) analysis was performed on an ESCALAB XI+ photoelectron spectrometer (Thermo Fisher Scientific, Waltham, MA, USA), utilizing a Mg Kα source (hv = 1253.6 eV). The binding energy scale was calibrated by referencing the C 1s peak at 284.8 eV. The microscopic morphology of the samples was observed by scanning electron microscopy (SEM, TESCAN MIRA LMS, Brno, Czech Republic), with an energy-dispersive spectrometer (EDS, TESCAN MIRA LMS, Brno, Czech Republic) integrated for elemental analysis. Electron paramagnetic resonance (EPR) spectroscopy was performed using an EPR200-Plus spectrometer (CIQTEK, Hefei, China) to determine the generation of ¹O₂ during the reaction. The test conditions were as follows: center field 3380 G, scanning range 60 G, microwave frequency 9.5 GHz, modulation frequency 100 kHz, and capture agent concentration 400 mM.

3.4. Degradation Tests and Analytical Methods

Degradation experiments were performed under dark conditions. A total of 10 mg of MnO₂/CN@SiO₂ catalyst was ultrasonically dispersed in 40 mL of 50 mg/L LEV solution. Subsequently, the solution was poured into a 50 mL quartz bottle with a lid, and the reaction temperature was controlled by a DC-0506 low-temperature constant temperature bath (Perfect Light, Beijing, China). After stirring for 10 min to reach adsorption equilibrium, 400 μL of 0.1 M PMS solution was added, resulting in a final 1 mM PMS concentration in the reaction mixture. The reaction solution was filtered several times through a 0.45 μm PTFE filter membrane. Then, samples were taken every 10 min, and the absorbance of each sample was measured at a wavelength of 287 nm.

The concentration of LEV was determined using a UV spectrophotometer (UV-8000, Metash, Shanghai, China) with a maximum absorption wavelength of 287 nm. The concentration of PMS in the solution was measured using the iodometric method. Specifically, 0.02 g of NaHCO₃ was dissolved in 2.8 mL of KI solution, followed by the addition of 0.2 mL of PMS. The absorbance of PMS was measured at 319 nm.

LEV degradation efficiency was calculated according to Equation (10). A proposed first-order kinetic model was used to simulate the LEV degradation reaction kinetics (Equation (11)). k_app is the apparent reaction constant, and C₀ and C_t represent the initial mass concentration of LEV and the mass concentration at a reaction time of t, respectively.

$$\mathrm{D}=[1-({\mathrm{C}}_{\mathrm{t}}/{\mathrm{C}}_0) \times 100\%]$$

(10)

$${\mathrm{k}}_{\mathrm{app}}=-\ln ({\mathrm{C}}_{\mathrm{t}}/{\mathrm{C}}_0)/\mathrm{t}$$

(11)

4. Conclusions

In summary, MnO₂/CN@SiO₂ catalysts were successfully synthesized, and their catalytic performance for activating PMS to degrade LEV was systematically investigated. The catalyst, prepared with the mass ratio of 5:4:1 for MnO₂:CN:SiO₂, exhibited the highest LEV removal of 80.8%, which was higher than those of MnO₂ (67.4%), MnO₂/CN (68.2%), and MnO₂/SiO₂ (76.1%). The incorporation of SiO₂ significantly enhanced the catalytic activity, attributed to the rapid adsorption of PMS onto the catalyst surface via Si–OH groups, facilitating its reaction with MnO₂. The experimental results revealed that the MnO₂/CN@SiO₂ catalysts exhibited excellent tolerance to low concentrations of ${\mathrm{NO}}_3^{-}$, ${\mathrm{SO}}_4^{2-}$, Cl⁻, and HA, and high stability in cycling experiments. Based on the free radical quenching experiment and EPR spectroscopy analysis, a non-radical-based pathway for LEV degradation was proposed. The primary degradation mechanism involved electron transfer from LEV to the Mn³⁺–(O)${\mathrm{O}\mathrm{S}\mathrm{O}}_3^{-}$ complex, which oxidizes LEV with high activity. The ¹O₂-mediated reaction also contributes to LEV degradation. In conclusion, the MnO₂/CN@SiO₂ catalyst demonstrates superior performance in activating PMS and holds significant potential for application in treating antibiotic-contaminated wastewater.