Mn₃O₄/biochar catalyst for enhanced PMS activation and antibiotic degradation via radical and non-radical pathways

Abstract

Metronidazole (MNZ) presents a significant threat to both human health and aquatic environments due to its environmental stability, antimicrobial properties, and capacity to promote drug resistance. Despite this, the effective and environmentally sound elimination of this compound remains a significant hurdle. This study introduces a hydrothermal and co-precipitation approach to synthesize a Mn₃O₄/biochar composite catalyst. This catalyst was engineered to activate peroxymonosulfate (PMS) and thereby expedite the removal of MNZ. The Mn₃O₄/biochar/PMS system exhibited outstanding catalytic efficacy, achieving a complete removal and a kinetic constant of 0.1402 min⁻¹ within a 30 minute timeframe under optimal conditions. The degradation of MNZ was consistently effective under various conditions, including different catalyst amounts, concentrations of PMS, initial pollutant concentrations, initial pH levels, the presence of inorganic anions, humic acid, and different water sources. The identification of reactive species through quenching experiments corroborated that sulfate (SO₄˙⁻) and hydroxyl radicals (˙OH) were the main contributing factors in activating PMS for MNZ degradation, along with the minor contribution of singlet oxygen (¹O₂). Mechanistic studies also revealed that the efficient activation of PMS was facilitated by redox cycling between Mn²⁺/Mn³⁺/Mn⁴⁺. Additionally, possible MNZ degradation pathways were identified through UPLC-QTOF/MS analysis. Consequently, this study accomplished the development of highly stable, environmentally friendly, and efficient catalysts for PMS activation with lower metal leaching, thereby offering a promising method for effectively removing persistent organic pollutants from real-world water sources.

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1. Introduction

Antibiotic residues in water systems are a global public health issue that has emerged in recent years, as they introduce the risk of the development of antibiotic resistance among microbes and possibly lead to serious health hazards.^1–3 The global health crisis is worsened by the indiscriminate discharge of agricultural and industrial pollutants, such as pesticides, organic pigments, and pharmaceutical byproducts, into water. Metronidazole (MNZ), a prominent nitroimidazole antibiotic, possesses both antibacterial and anti-inflammatory properties.^4,5 Owing to its high water solubility and lower molecular weight, MNZ can penetrate microbial cell membranes, where it is reduced, potentially resulting in damage to microbial DNA.⁶ The chemical structure of MNZ features a five-membered heterocyclic ring linked to a nitro group that contributes to the structure's stability, thereby complicating its removal via standard water treatment processes.⁷ As a result, the environment has been subject to considerable MNZ contamination, which negatively affects the health of both humans and animals by promoting microbial development and reproduction.^8,9 Consequently, it is imperative to devise a viable approach for the degradation of the MNZ antibiotic in the aqueous environment.¹⁰

In recent years, there has been significant research interest in electrocatalysis, photocatalysis, heterogeneous catalysis, adsorption, membrane filtration and advanced oxidation processes for elimination of toxic organic pollutants from aqueous solutions.^11,12 Advanced oxidation processes utilizing peroxymonosulfate (PMS) activation can be designed to degrade antibiotics via the generation of superoxide (O₂˙⁻), hydroxyl (˙OH), and sulfate (SO₄˙⁻) radicals, as well as non-radical singlet oxygen (¹O₂).^13,14 The SO₄˙⁻ radical demonstrates a broader pH range (2–9) responsiveness, and a larger reduction potential (2.5–3.0 V) in contrast to the ˙OH radical.¹⁵ Various procedures have been employed to induce PMS to generate SO₄˙⁻, including UV light, heat, ultrasonication, carbonaceous materials, and transition metals.¹⁶

Practical implementations in environmental areas necessitate the advancement of strong catalysts for the elimination of antibiotics. Nevertheless, the aggregation of catalyst material resulted in a reduction in its surface area, which in turn impeded catalyst activity and increased energy consumption and PMS.¹⁷ Recently, the utilization of transition metal-based catalysts to activate PMS is considered a promising approach.^18,19 Their use is mainly preferred due to their operational simplicity, low energy needs, the reduction of pollutants, and their widespread availability.²⁰ However, the leaching of metal ions during the catalytic process is a common problem for certain catalysts, such as manganese, iron, nickel, and cobalt.²¹ The process of leaching can lead to secondary pollution and potential health risks, which limits the practical use of these materials in water treatment.²² Therefore, creating catalysts that are structurally stable, maintain high activity for a long time, and have a minimal environmental impact remains a significant challenge.

Manganese-based materials, particularly manganese oxides (Mn₃O₄), are highly valued among transition metal-based activators because they are abundant, eco-friendly, low-cost, have mixed-valence chemistry, and are stable in the environment.²³ Mn₃O₄ is known for its inherent stability, which is a result of its unique distorted spinel structure. This structure consists of manganese ions in two different oxidation states, Mn²⁺ species that occupy the tetrahedral sites, and Mn³⁺ species that occupy octahedral sites.²⁴ Such a long-range ordered distribution of Mn²⁺ and Mn³⁺ ions could be beneficial for PMS activation.²⁵ So, the Mn₃O₄/PMS system has shown significant efficiency in the degradation of organic pollutants.²⁶ Despite showing exceptional PMS activation catalytic activity, the agglomeration of Mn₃O₄ nanoparticles due to their poor thermal and chemical stabilities hindered reproducibility.²⁷ Simultaneously, the retrieval of the particulate catalyst without the presence of a carrier is not only a complex process, but it is also difficult to avoid the continuous leaching of the nanoparticles into the environment.²⁸ Carbonaceous materials have been extensively employed as the matrix for Mn₃O₄-based materials to enhance their stability and conductivity in order to address these obstacles.²⁹

Biochar, a carbonaceous porous substance, presents an attractive combination of affordability, sustainability via waste valorization, elevated specific surface area, and easily alterable surface chemistry.^30–32 Biochar is formed through the pyrolysis of biomass under oxygen-limited or anoxic conditions at high temperatures.³³ It can be widely employed as a catalyst support due to its environmental friendliness and benefit for the PMS activation process.³⁴ Nonetheless, the catalytic efficacy of unmodified biochar often proves inadequate for PMS activation due to its chemical inertness.³⁵ Therefore, initiatives to improve its performance are being aggressively implemented to realize its full potential as an effective catalyst. Combining transition metal oxides with biochar has shown promise in creating complex structures with many active sites, which could help activate PMS.³⁶

Based on the exceptional catalytic properties of Mn₃O₄ and biochar as a supporting material, we have developed highly efficient Mn₃O₄/biochar through hydrothermal and co-precipitation methods. Mn₃O₄/biochar was used to trigger PMS and, thus, exploited for the elimination of the MNZ antibiotic from wastewater. The surface morphology, crystallinity, surface chemistry, and other features of the produced Mn₃O₄/biochar catalyst were determined through a diverse characterization technique. Mn₃O₄/biochar's catalytic competence was further examined by MNZ degradation under varied circumstances. Additionally, the mitigation of MNZ in the presence of ionic species, naturally occurring organic matter, and various water systems using Mn₃O₄/biochar catalyst was examined. The identification of ROS was accomplished through radical scavenging experiments and a plausible degradation method was deduced based on the intermediate products. Finally, the interaction between the Mn₃O₄/biochar and PMS along with the elimination of contaminants has been lavishly explained by this study.

2. Experimental

2.1. Chemicals

KMnO₄, polyethylene glycol (PEG), sodium hydroxide (NaOH), potassium peroxymonosulfate (2KHSO₅ KHSO₄ K₂SO₄), metronidazole (C₆H₉N₃O₃), HNO₃, NaCl, NaNO₃, NaHCO₃, humic acid, CH₃OH, C₂H₅OH, and tert-butyl alcohol (C₄H₁₀O), L-histidine (C₆H₉N₃O₂), were purchased from Sigma Aldrich. Throughout the experiments, deionized water was used for sample preparation and catalytic activity evaluation.

2.2. Synthesis of Mn₃O₄/biochar

In this investigation, wheat straw was chosen as the standard lignocellulosic biomass due to its cost-effectiveness and global availability for biochar production. The biochar was synthesized through the pyrolysis of wheat straw in a tubular furnace at 500 °C with a heating rate of 5 °C min⁻¹ for 180 minutes under a nitrogen atmosphere.

For the preparation of Mn₃O₄, KMnO₄ (0.79 g) was solubilized in water (30 mL) under continuous agitation to homogenize the mixture. Then, introduce 30 mL of PEG to the KMnO₄ solution and continue agitation for an additional 30 minutes. The processed mixture was taken in a 100 mL Teflon-lined autoclave which was maintained at 120 °C for 10 h. When cooled, the reddish-brown Mn₃O₄ was vacuum filtered and washed several times alternately with ethanol and deionized water. Ultimately, the powdered material was heated to 50 °C and desiccated in a dehydrating oven for 10 hours. The final sample was then named Mn₃O₄.

Subsequently, the Mn₃O₄/biochar composite was synthesized via a simple chemical coprecipitation method. Initially, biochar (0.5 g) was dispersed in C₂H₅OH (50 mL) for a duration of five minutes to form suspension A. Concurrently, suspension B was obtained by dispersing Mn₃O₄ (1 g) powder in water (50 mL) for one hour. Following this, suspensions A and B were amalgamated and subjected to magnetic stirring for a period of two hours. Following vacuum filtering, the as-synthesized Mn₃O₄/biochar composites were dried at 70 °C for half a day. The mass ratio of Mn₃O₄ to biochar for the synthesis of Mn₃O₄/biochar was 2 : 1 (1 g Mn₃O₄ and 0.5 g biochar), corresponding to about 33 wt% biochar in the composite.

2.3. Catalytic evaluation of Mn₃O₄/biochar

To analyze the catalytic performance of prepared catalysts, MNZ, a common antibiotic, was considered as a probe, and the degradation was performed using PMS. The experimental process was accomplished in a flask, containing 100 mL of MNZ solution (30 mg L⁻¹) and a catalyst concentration of 0.45 g L⁻¹ at room temperature and with continuous stirring. Later on, the reaction mixture was kept for ½ h to facilitate sorption equilibrium. The degradation process of MNZ was initiated by incorporating the PMS (0.24 mM) to the reaction mixture. During degradation reactions, the pH was adjusted by introducing 1.0 M solutions of NaOH and HNO₃. After taking 0.5 mL samples of the mixture at different times during the degradation, the reaction was stopped by adding an equal amount of methanol. The resulting mixture was subsequently filtered using PTFE filters (pores = 0.22 µm). HPLC was implemented to measure the residual MNZ. The degradation of MNZ was affected by several factors: the amount of PMS used (0, 0.08, 0.16, 0.24, and 0.32 mM), the concentration of the catalyst (0.0, 0.15, 0.30, 0.45, and 0.60 g L⁻¹), the initial concentration of MNZ (10, 20, 30, 40, and 50 mg L⁻¹), and the pH of the solution (3, 4.86 (unadjusted), 7, 9, and 11). In addition, the effects of ionic species and natural organic molecules on MNZ decomposition were studied to assess the catalyst's potential for treating real wastewater. Furthermore, radical quenching studies were carried out to determine the major reactive oxygen species (ROS) in MNZ degradation. Scavengers were introduced at the beginning of the process with the concentrations as follows: methanol (100 mM) and tert-butyl alcohol (TBA, 100 mM) for ˙OH/SO₄˙⁻ and ˙OH respectively, while L-histidine (10 mM) for ¹O₂.

3. Results and discussion

3.1. Catalyst characterizations

As illustrated in Fig. 1, XRD study was conducted to investigate the lattice structure and crystal phase purity of Mn₃O₄, biochar, and Mn₃O₄/biochar. From Fig. 1, it is clear that pristine Mn₃O₄ unveiled distinctive diffraction peaks at 18.08°, 28.96°, 31.18°, 32.42°, 36.21°, 38.21°, 44.65°, 50.82°, 53.96°, 56.17°, 58.60°, 60.07°, and 64.88°. The observed peaks correspond to the (101), (112), (200), (103), (211), (004), (220), (105), (312), (303), (321), (224), and (400) planes, respectively, confirming the formation of highly crystalline tetragonal hausmannite Mn₃O₄ (JCPDS no. 24–0734) with an I4₁/amd space group characterized by mixed Mn²⁺/Mn³⁺ valence states.^37,38 On the other hand, the diffraction peaks ranging from 21.83° to 24.3° in the XRD spectra of biochar were related to the (002) plane of amorphous carbon,³⁹ while the peak at 26.58° was related to the (111) plane of graphitic-carbon.⁴⁰ The successful formation of Mn₃O₄/biochar composites is confirmed by the presence of all characteristic planes of both Mn₃O₄ and biochar in the XRD spectra. However, the peak intensity of biochar in Mn₃O₄/biochar composites exhibited a marked reduction when compared to the pure biochar, which can be attributed to their comparatively low concentration within the composite structure.

Fig. 1 XRD spectrum of Mn₃O₄, biochar, and Mn₃O₄/biochar.

SEM was used to examine the shape and structure of Mn₃O₄/biochar. The SEM micrographs in Fig. 2(a–c) illustrate the surface topography of the synthesized Mn₃O₄/biochar composite at different magnifications. The Mn₃O₄ phase shows anisotropic crystal growth and is characterized by small and particle-like crystals that are combined with biochar. The biochar surface has effectively immobilized the heterogeneous distribution of Mn₃O₄ particles, resulting in a multi-phase composite of a carbon matrix and metal oxide particles. Furthermore, SEM-EDS elemental mapping analysis verifies the effective formation of the Mn₃O₄/biochar composite. Fig. 2(d–g) indicates that the composite surface was distributed with C, O, and Mn. This suggests that Mn₃O₄ was deposited onto the biochar material.

Fig. 2 (a–c) SEM micrographs of Mn₃O₄/biochar at different magnifications, (d–g) EDS mapping and (h) EDS quantitative analysis (inset representing the weight percentages) of Mn₃O₄/biochar.

The elemental makeup is substantiated by the EDS spectrum (Fig. 2(h)), which displays prominent peaks associated with C, O, and Mn, specifically with weight percentages of 26.4% for C, 20.4% for O, and 53.2% for Mn. The elevated concentrations of Mn and O, alongside a suitable C content, suggest that Mn₃O₄ particles are effectively anchored to the biochar surface, rather than being present as a distinct phase. As a result, the consistent distribution of all three elements, absent of substantial impurities, confirms the successful fabrication of the Mn₃O₄/biochar composite.

The surface chemical makeup and the changes in valence states of Mn₃O₄/biochar were interpreted through XPS study. The comprehensive scan spectrum of Mn₃O₄/biochar shown in Fig. 3(a)) validated the existence of O, Mn, and C in the composite sample. The Mn 2p spectrum (Fig. 3(b)) displayed two main peaks, Mn 2p_3/2 and Mn 2p_1/2, at 641.84 eV and 653.42 eV. The difference in spin–orbit binding energy between these peaks was found to be 11.58 eV, which is consistent with previous research on Mn₃O₄. (ref. 41) In addition, the Mn 2p_3/2 peak could be divided into distinct spikes at 641.02, 642.24, and 643.81 eV, related to Mn²⁺, Mn³⁺, and Mn⁴⁺, separately.⁴² The O 1s spectra of the Mn₃O₄/biochar catalyst shown in Fig. 3(c) and its deconvolution yielded three spikes at 529.62, 531.16, and 532.71 eV, which are ascribed to lattice oxygen, –OH, and adsorbed H₂O species, respectively.⁴³ The –OH species on the catalyst surface serve as active sites for bond formation with PMS.⁴⁴ The C 1s spectra (Fig. 3(d)) showed peaks at 284.36, 285.45, and 288.27 eV. These peaks were accompanied by C=C/C–C, C–O, and C–O–C functional groups, respectively.⁴⁵

Fig. 3 XPS spectral analysis of Mn₃O₄/biochar; (a) survey scan, (b) Mn 2p, (c) O 1s, (d) C 1s.

The active sites' population on a catalyst's surface is strongly related to its surface area. Fig. 4(a and b) illustrates the adsorption–desorption curves of N₂ for bare Mn₃O₄ and Mn₃O₄/biochar. The BET analysis of both Mn₃O₄ and Mn₃O₄/biochar revealed type IV isotherms, which implies that both materials were typical mesoporous substances.^46,47 The Mn₃O₄/biochar composite established a larger surface area (67.82 m² g⁻¹) in contrast to the bare Mn₃O₄ (45.29 m² g⁻¹), thus offering a greater density of active sites and facilitating the removal of organic pollutants. The observed increase in the surface area of Mn₃O₄/biochar validated the successful integration of Mn₃O₄ with biochar to form a composite material. The Mn₃O₄/biochar composite is expected to speed up the degradation process and help in the adsorption of pollutant molecules due to its mesoporous structure and larger surface area.

Fig. 4 N₂ adsorption/desorption curves of (a) Mn₃O₄ and (b) Mn₃O₄/biochar.

3.2. Performance of Mn₃O₄/biochar for the degradation of MNZ antibiotic

The adsorption removal of MNZ was evaluated to study an interesting relationship between the adsorption performance and catalytic efficiency, as shown in Fig. 5(a). The results indicate that the adsorptive removal of MNZ was quite limited in the presence of various catalysts and reached 7.92% for Mn₃O₄/biochar/PMS, 6.71% for Mn₃O₄/biochar, 5.36% for Mn₃O₄/PMS, 4.32% for Mn₃O₄, and 2.92% for biochar after 30 min of adsorption. Furthermore, the degradation potential of the synthesized Mn₃O₄, biochar, and Mn₃O₄/biochar catalysts was gauged with and without PMS (Fig. 5(b)). Without the presence of any catalyst, the effective removal of MNZ presents a significant challenge, as demonstrated by the negligible (2.65%) reduction in its concentration observed over the 30 minute reaction period, implying that PMS alone could not degrade MNZ. Similarly, the degradation of MNZ was limited to 5.39%, 13.89%, and 23.24% during the same period when biochar, Mn₃O₄, and Mn₃O₄/biochar were utilized in the absence of PMS. The Mn₃O₄/PMS system, however, significantly enhanced the elimination of MNZ and led to the abatement of 74.09% of MNZ in a 30 minute reaction. On the other hand, the utilization of Mn₃O₄/biochar/PMS system as an efficient PMS activator markedly promoted the MNZ elimination, with a complete degradation, within 30 minutes. It was due to the fact that the transition of Mn²⁺/Mn³⁺/Mn⁴⁺ resulted in the activation of PMS to yield highly ROS for the abatement of MNZ.

Fig. 5 a) MNZ adsorption in the presence of various reaction systems, (b) MNZ degradation rates in the presence of various reaction systems, (c) study of kinetics, (d) corresponding k_app, and (e) Mn ions leaching in Mn₃O₄/PMS and Mn₃O₄/biochar/PMS systems. (Parameters: catalyst = 0.45 g L; PMS = 0.24 mM; MNZ = 30 mg L; pH = 4.86).

The kinetic investigation of the specified catalytic systems for MNZ degradation was performed employing the Langmuir–Hinshelwood model, as represented by eqn (1).⁴⁸

Here, k_app signifies the apparent rate constant, while C₀ and C refer to the MNZ initial and final concentrations at time t, respectively. As depicted in Fig. 5(c), the various catalyst systems for MNZ removal exhibited pseudo-first-order kinetics. Fig. 5(d) presents the k_app values for MNZ degradation, which were determined to be 0.1402, 0.047, 0.0084, 0.005, and 0.0016 min⁻¹ for the Mn₃O₄/biochar/PMS, Mn₃O₄/PMS, Mn₃O₄/biochar, Mn₃O₄, and biochar, respectively. The k_app of the Mn₃O₄/biochar/PMS system was 2.98 times larger than Mn₃O₄/PMS, indicating that the Mn₃O₄/biochar/PMS system demonstrated a superior degrading capacity. The enhanced MNZ degradation is attributable to the beneficial characteristics and augmented surface area of Mn₃O₄/biochar in facilitating PMS activation. Furthermore, to highlight the importance of our synthesized catalyst, a comparison was made between the catalytic effectiveness of Mn₃O₄/biochar and that of previously researched catalysts (Table 1).

Table 1 Catalytic superiority of Mn₃O₄/biochar over previously reported catalysts for PMS activation towards the degradation of MNZ antibiotics

Catalyst	Synthesis	Pollutant	Degradation	Reaction time	Ref.
NSC-Co3O4	Solvothermal/pyrolysis	MNZ	99%	30 min	49
Fe-Ce@N-BC	Hydrothermal	MNZ	97.5%	60 min	50
NiFePx	Phosphorization	MNZ	94.5%	60 min	51
FeCo2O4-Fe3O4	Sol–gel	MNZ	96.8%	60 min	52
Co-MBC	Impregnation pyrolysis	MNZ	90%	60 min	53
LaCo0.8Mn0.2O3	Sol–gel	MNZ	99.42%	30 min	54
C-MCED/Fe(III)	—	MNZ	94.05%	210 min	55
CoAl2O4@AP	In situ growth/calcination	MNZ	97%	100 min	56
Fe/Mn–BC	—	MNZ	94.3%	120 min	57
Mn3O4/biochar	Hydrothermal/co-precipitation	MNZ	100%	30 min	This study

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The release of Mn ions was assessed in leaching experiments, which were performed under optimal reaction conditions to determine the practical utility of Mn₃O₄ and Mn₃O₄/biochar catalysts during PMS activation (Fig. 5(e)). The concentrations of leaked Mn ions were measured, yielding values of 0.627 and 0.404 mg L⁻¹ for Mn₃O₄/PMS and Mn₃O₄/biochar/PMS systems, respectively. These values are significantly lower than the permitted discharge limit (GB 3838–2002, 1.0 mg L⁻¹),⁵⁸ indicating that the synthesized catalysts are suitable for their intended use. Additionally, the leaching results verified that the improved efficiency of Mn₃O₄/biochar was a consequence of the composite formation, which led to a decrease in metal ion leaching.

3.3. Optimization of interfering factors

3.3.1. Influence of Mn₃O₄/biochar loading. To understand the effectiveness of MNZ removal using the Mn₃O₄/biochar/PMS system under the influence of various catalyst loadings, we did degradation experiments with different amounts of catalyst in the range of 0 to 0.60 g L⁻¹. The elimination of MNZ enlarged from 57.62% to complete degradation as the catalyst loading was increased from 0.15 to 0.45 g L⁻¹ (Fig. 6(a)). The k_app for MNZ elimination was around 0.0295, 0.0716, and 0.1402 min⁻¹ at catalyst loadings of 0.15, 0.30, and 0.45, respectively (Fig. 6(b)). The enhanced MNZ abatement efficiency at 0.45 g L⁻¹ resulted from the accessibility of supplementary active sites for PMS stimulation, which contributed to the rise in mitigation rate as catalyst loading was raised.⁵⁹ Conversely, the MNZ mineralization efficacy experienced a slight decrease, reaching 97.07%, at a catalyst dosage of 0.60 g L⁻¹. Furthermore, the k_app value was correspondingly reduced to 0.1188 min⁻¹. The quenching of free radicals may be attributed to the recombination of surplus radicals or the self-aggregation of excess catalysts, which could explain the decrease in degradation efficacy.⁶⁰

Fig. 6 MNZ degradation under the influence of (a and b) Mn₃O₄/biochar loading, (c and d) PMS quantity, (e and f) MNZ concentration, and (g and h) pH using the Mn₃O₄/biochar/PMS system.

3.3.2. Influence of PMS quantity. To determine the influence of PMS conc. on the effectiveness of MNZ elimination in the Mn₃O₄/biochar/PMS system, trials were performed using different PMS amounts (Fig. 6(c)). Evidently, increasing the con. of PMS markedly improved the effectiveness of MNZ removal. The MNZ abatement proficiency was substantially enhanced from 48.16% to complete degradation as the PMS dosage enlarged from 0.08 to 0.24 mM, while the k_app enlarged from 0.0215 to 0.1402 min⁻¹ (Fig. 6(d)). By contrast, a slight reduction in removal efficiency to 95.58% was recorded, with a k_app of 0.1038 min⁻¹, upon increasing the PMS concentration to 0.32 mM. These observations can be explained by two key factors: (i) the Mn₃O₄/biochar catalyst demonstrated an enhanced capacity to activate PMS, thereby generating a greater quantity of ROS, which, in turn, accelerated the elimination of MNZ when the PMS concentration was at 0.24 mM. (ii) Simultaneously, the surplus PMS may have interacted with the excess SO₄˙⁻ radical,⁶¹ consequently producing less ROS, causing a decline in decomposition efficacy as the PMS concentration surpassed 0.32 mM. Subsequently, the reasonable quantity of PMS for successive trials was determined to be 0.24 mM.

3.3.3. Influence of MNZ concentration. The consequence of the MNZ quantity on the degradation process was investigated, and the obtained results for different MNZ con (Fig. 6(e)). The degrading efficacy was adversely affected by the increase in MNZ initial concentration. In fact, the abatement of MNZ was more facilitated by lower initial concentrations. At 10 mg L⁻¹, MNZ can be completely degraded in 25 min, while at 20 and 30 mg L⁻¹, complete degradation is achieved in 30 min. Similarly, the observed k_app values for MNZ degradation decreased significantly, from 0.1598 min⁻¹ to 0.0825 min⁻¹, as the MNZ amount amplified from 10 to 50 mg L⁻¹ (Fig. 6(f)). The explanation may be further emphasized by the fact that the number of ROS was restricted as a result of the limited availability of PMS and catalyst.⁶² Therefore, the degradation efficacy may be reduced as a result of the competition for ROS in the reaction system, which is caused by the increase in MNZ molecules.

3.3.4. Influence of the pH. The influence of the pH of the solution is another critical aspect that alters the surface belongings of the catalyst and controls the active species generation.⁶³ Therefore, the trials were performed over a pH assortment of 3–11 to investigate the impact of pH on the decomposition of MNZ using the Mn₃O₄/biochar/PMS, as shown in Fig. 6(g). The outcomes indicated that MNZ mitigation was more efficient under slightly acidic and neutral conditions. The Mn₃O₄/biochar/PMS system achieved complete MNZ removal at pH 4.86 (unadjusted initial pH), thereby showcasing its high MNZ degradation capability. Furthermore, even at pH 7, MNZ degradation was still substantial, reaching 95.57%, which highlights the catalyst's broad applicability across a range of pH levels. Conversely, MNZ degradation was limited to 87.85%, 86.59%, and 83.72% at pH 3, 9, and 11, respectively. The k_app for pH 3, 4.86, 7, 9, and 11 were 0.0733, 0.1402, 0.1032, 0.0685, and 0.0613 min⁻¹, separately (Fig. 6(h)). At pH 4.86, the elevated k_app value of 0.1402 min⁻¹ correlates with the most significant degree of MNZ elimination. Because of the rise in H⁺ concentration at pH 3, the interface among PMS and MNZ, which may boost the establishment of H-bonding via the H⁺ and HSO₅⁻, is repressed under very acidic circumstances.⁶⁴ Additionally, as seen in eqn (2) and (3), H⁺ may reduce the effectiveness of MNZ degradation by quenching SO₄˙⁻ and ˙OH.

H⁺ + SO₄˙⁻ + e⁻ → HSO₄˙⁻ (2)

H⁺ + ˙OH + e⁻ → H₂O (3)

It is also important to note that in basic conditions, the reaction among SO₄˙⁻ and H₂O/⁻OH might produce ˙OH (eqn (4) and (5)), which has a reduced redox potential, causing the degradation ineffectiveness.

SO₄˙⁻ + H₂O → SO₄²⁻ + ˙OH + H⁺ (4)

SO₄˙⁻ + –OH → SO₄²⁻ + ˙OH (5)

3.3.5. Influence of coexisting anions. A variety of inorganic ions and HA have been demonstrated to effectively quench the ROS generated through the PMS activation, and the incidence of such specific ions in water may impact the efficiency of the Mn₃O₄/biochar. Numerous inorganic ions have been seen to rapidly quench the reactive radicals. The presence of such ions may significantly influence the catalyst's ability. The impact of these ions on the MNZ abatement via Mn₃O₄/biochar catalyst is seen in Fig. 7. The data in Fig. 7(a) demonstrate that 10 mM Cl⁻ somewhat obstructs the removal rate of MNZ, reducing it to 92.41%. Cl⁻ can quickly undergo a reaction with SO₄˙⁻ or ˙OH radicals produced by the Mn₃O₄/biochar catalyst through PMS activation. This reaction leads to the generation of chlorinated radicals (Cl₂˙⁻), which have a lower oxidation potential, thereby decreasing the degradation efficiency of MNZ (eqn (6)–(9)).⁶⁵

SO₄˙⁻ + Cl⁻ → SO₄²⁻ + Cl˙ (6)

Cl⁻ + ˙OH → ClOH˙ (7)

ClOH˙ + H⁺ → Cl˙ + H₂O (8)

Cl˙ + Cl⁻ → Cl₂˙⁻ (9)

Fig. 7 MNZ degradation under the influence of (a) co-existing anions and HA and (b) different water systems using the Mn₃O₄/biochar/PMS system. (Parameters: Mn₃O₄/biochar = 0.45 g L; PMS = 0.24 mM; MNZ = 30 mg L; pH = 4.86).

Conversely, the introduction of 10 mM NO₃⁻ only slightly slows down the breakdown of MNZ, reaching 95.67%. This effect may be due to the reaction between NO₃⁻ and SO₄˙⁻, which produces less reactive radicals (eqn (10)), thereby hindering the degradation process.

Conversely, the presence of HCO₃⁻ significantly reduced the degradation of MNZ. Specifically, adding 10 mM HCO₃⁻ repressed the reduction of MNZ, resulting in a total degradation of 84.52%. This decrease in abatement was ascribed to the interaction between HCO₃⁻ and SO₄˙⁻/˙OH, which led to the generation of , a species with a lower oxidation potential, thereby hindering effective MNZ degradation (eqn (11) and (12)).⁶⁶

HCO₃⁻ + ˙OH → CO₃˙⁻ + H₂O (12)

Conversely, the presence of HPO₄²⁻ significantly reduced the effectiveness of MNZ removal to 90.64%. HPO₄²⁻ may react with SO₄˙⁻/˙OH, hence hindering MNZ degradation (eqn (13) and (14).⁶⁷ Furthermore, the crystallization of HPO₄²⁻ with metallic ions could reduce the catalyst's effectiveness in activating PMS.

HPO₄²⁻ + SO₄˙⁻ → HPO₄˙⁻ + SO₄²⁻ (13)

HPO₄²⁻ + ˙OH → HPO₄˙⁻ + OH⁻ (14)

Moreover, HA, a common redox-active substance, has the potential to hinder MNZ reduction. This occurs through the blockage of the catalyst's active sites and competition for the production of ROS. In accordance with this premise, the MNZ mineralization rate was assessed by adding 20 mg L⁻¹ of HA, and the results are presented in Fig. 7(a). The mitigation of MNZ was shown to be a diminution from complete to 89.43% over a 30 minute reaction, thereby indicating that even a modest concentration of HA can substantially hinder MNZ degradation.

To assess the practical utility of the Mn₃O₄/biochar/PMS system in real-world aquatic environments, tap water and artificial lake water served as representative matrices for investigating MNZ removal across diverse water quality conditions. In tap water and lake water, the removal efficiencies of MNZ, as shown in Fig. 7(b), were 94.11% and 87.66%, respectively. This suggests that the system effectively breaks down MNZ under conditions similar to those found in real-world situations. These findings reveal that the Mn₃O₄/biochar/PMS system has definite aptitude for real-world use in wastewater treatment.

3.4. Identification of active species

To identify the predominant ROS produced during the breakdown of MNZ using the Mn₃O₄/biochar/PMS, radical trapping studies were conducted as seen in Fig. 8(a and b). In most PMS-based AOPs, SO₄^•−/^•OH radicals are the main ROS that have been identified. Methanol served as an active quenching agent for SO₄˙⁻/˙OH radicals. Conversely, tert-butyl alcohol (TBA) quickly reacts with ˙OH radicals.⁶⁸ Therefore, methanol and TBA were used as specific indicators for SO₄˙⁻/˙OH and ˙OH, separately.^69,70 Upon the addition of methanol, the MNZ elimination rate reduced to 44.61%, yielding k_app of 0.0193 min⁻¹. Moreover, using the TBA, the degrading efficiency decreased to 55.37%, accompanied by a k_app of 0.0272 min⁻¹. These findings demonstrated that SO₄˙⁻ and ˙OH were generated in the Mn₃O₄/biochar/PMS system, with SO₄˙⁻ being one of the most predominant radicals. Alongside SO₄˙⁻ and ˙OH, other prevalent non-radicals, including ¹O₂, also played a role in the breakdown of MNZ. Consequently, L-histidine was chosen as a scavenger for ¹O₂. The removal efficiencies of MNZ, as shown in Fig. 8(a and b), decreased to 91.04% with the addition of L-histidine. This was accompanied by a reduction in k_app values to 0.0805 min⁻¹. Therefore, the results from the quenching experiments suggested that SO₄˙⁻/˙OH/¹O₂ were the main ROS participating in MNZ removal within the Mn₃O₄/biochar/PMS system.

Fig. 8 (a) Radical scavenging experiments for the determination of active species participating in the MNZ abatement using the Mn₃O₄/biochar/PMS system and (b) corresponding k_app values. (Parameters: Mn₃O₄/biochar = 0.45 g L; PMS = 0.24 mM; MNZ = 30 mg L; pH = 4.86).

3.5. Degradation mechanism and pathways

The degradation of MNZ within the Mn₃O₄/biochar/PMS system was postulated, as illustrated in Fig. 9, based on the preceding characterizations and experimental findings. Initially, Mn₃O₄/biochar and PMS were introduced into the MNZ solution. Subsequently, PMS was adsorbed onto the Mn₃O₄/biochar surface via electrostatic attraction, initiating a reaction that generated free radicals. Electrons from the oxidation processes of Mn²⁺/Mn³⁺ and Mn³⁺/Mn⁴⁺ produced SO₄˙⁻, which activated PMS (eqn (15) and (17)). The oxidation states of Mn (Mn³⁺/Mn⁴⁺) could be transformed into the reductive states of Mn (Mn²⁺/Mn³⁺). This resulted in the formation of SO₅˙⁻ (eqn (16) and (18)), which was important to enhance the activity of Mn₃O₄ and preserve its stability. In the meantime, the breakdown of MNZ was also linked to the formation of ˙OH by the interaction of SO₄˙⁻ with H₂O or –OH (eqn (19) and (20)). However, the created SO₅^•− would also contend with the –OH, leading to the production of ˙OH. Additionally, L-histidine quenching experiments verified the important role of ¹O₂ in the MNZ abatement. The SO₅˙⁻ radical produced may also combine with water or itself to yield ¹O₂ (eqn (21) and (22)). Ultimately, the generated SO₄˙⁻ and ˙OH radicals and ¹O₂ non-radicals can transform MNZ into low molecular weight non-toxic products (eqn (23)).

Mn²⁺ + HSO₅⁻ → Mn³⁺ + OH⁻ + SO₄˙⁻ (15)

Mn³⁺ + HSO₅⁻ → Mn²⁺ + H⁺ + SO₅˙⁻ (16)

Mn³⁺ + HSO₅⁻ → Mn⁴⁺ + OH⁻ + SO₄˙⁻ (17)

Mn⁴⁺ + HSO₅⁻ → Mn³⁺ + H⁺ + SO₅˙⁻ (18)

SO₄˙⁻ + H₂O → SO₄²⁻ + H⁺ + ˙OH (19)

SO₄˙⁻ + –OH → SO₄²⁻ + ˙OH (20)

2SO₅˙⁻ → 2SO₄²⁻ + ¹O₂ (21)

2SO₅˙⁻ + H₂O → 2HSO₄⁻ + 1.5¹O₂ (22)

SO₄˙⁻/˙OH/¹O₂ + MNZ → Degradation products (23)

Fig. 9 PMS activation and MNZ degradation mechanism using Mn₃O₄/biochar/PMS.

The intermediates of MNZ were examined using UPLC-Q-TOF-MS in order to get a better understanding of the degradation processes of MNZ in the Mn₃O₄/biochar/PMS system. During the degradation process, a total of eight degradation products were likely identified based on fragmentation structures. For the mitigation of MNZ using Mn₃O₄/biochar/PMS systems, there are two feasible degradation pathways based on fragmentation products (Fig. 10). The first pathway elaborates denitration and H-removal, that ultimately resulted in the production of a nitrogen–carbon-centered radical intermediate P1 (m/z = 125) due to the SO₄˙⁻ radical assault on the MNZ imidazole ring. Subsequently, P1 underwent cleavage at the N=C junctions, yielding a product designated P2 (m/z = 89). The transformation of P2 produced an olefinic intermediate, P3 (m/z = 88), which was then oxidized to create P4 (m/z = 90).⁷¹ Conversely, the second pathway commenced with the production of P5 (m/z = 186) and P6 (m/z = 127) via a hydroxyethyl cleavage reaction. Following this, P7 (m/z = 101) and P8 (m/z = 83) were produced through a series of nitro-reduction, N-denitration, and oxidation reactions.⁷² Finally, the straightforward intermediates P4 and P8 underwent further degradation into smaller organic molecules, such as acetic acid, prior to complete mineralization.

Fig. 10 Suggested MNZ degradation pathways using the Mn₃O₄/biochar/PMS system.

3.6. Catalyst stability

Regarding catalyst reusability and stability, five uninterrupted trials were executed under controlled settings to measure the catalyst's potential for repetitive use in large-scale, economically feasible applications. After every trial, the Mn₃O₄/biochar was separated from the reaction mixture through centrifugation. Then, it was carefully rinsed and then dried at 70 °C overnight to prepare it for reprocessing in the succeeding cycle. The efficiency of MNZ exhibited a gradual decrease owing to the repeated operation of the catalyst. Fig. 11(a) illustrates that the MNZ deterioration demonstrated a gradual reduction from complete degradation to 97.59%, 95.37%, 91.60%, and 87.45% during five consecutive cycles. The reduction in the degradation efficiency of Mn₃O₄/biochar involved several factors. First, partial leaching of active Mn species during reaction may decrease the number of accessible catalytic sites.⁷³ Second, the catalyst surface can be covered by the accumulation of chemical intermediates or by-products, which can obstruct the active sites and prevent the effective interaction between PMS and the catalyst. Third, the biochar support might undergo structural or surface changes, such pore blocking or surface oxidation, that might be responsible for the lower performance. Moreover, the repetitive use may lead to the agglomeration of Mn₃O₄ particles, which may further reduce the active surface area.

Fig. 11 (a) Reusability of Mn₃O₄/biochar/PMS system for the abatement of MNZ, (b) leaching amounts of Mn ions, and (c) XRD spectra of fresh and used (after five cycles) of Mn₃O₄/biochar catalyst.

Fig. 11(b) depicts the quantified leaching of Mn ions, resulting in values of 0.404, 0.32, 0.27, 0.23, and 0.21 mg L⁻¹ during five successive cycles. Furthermore, the XRD analysis of the used catalyst was performed to study the structural change after repeated use (Fig. 11(c)). The results suggest that the crystal structure of Mn₃O₄/biochar is mostly retained throughout reaction cycles. However, a small reduction in the peak intensity and small peak shifting was detected. These changes can be due to either partial structural distortion, surface modification or leaching of active Mn species to a small extent during the reaction phase. Finally, the aforementioned findings have validated the durability and reusability of Mn₃O₄/biochar catalyst, enabling its utilization in several degradation trials for MNZ abatement.

4. Conclusions

The present investigation established the efficacy of a Mn₃O₄/biochar catalyst, synthesized via a hydrothermal and co-precipitation method, in facilitating the activation of peroxymonosulfate (PMS) for the degradation of metronidazole (MNZ). Characterization of the catalysts revealed that Mn₃O₄/biochar possessed a more substantial surface area of 67.82 m² g⁻¹, whereas pure Mn₃O₄ presented a surface area of 45.29 m² g⁻¹. This disparity facilitated an increased availability of active sites for the triggering of PMS during the elimination of MNZ. Under ideal conditions, the MNZ attained complete degradation in 30 minutes having a k_app of 0.1402 min⁻¹. Conversely, the presence of inorganic ions, including Cl⁻, NO₃⁻, HCO₃⁻, and HPO₄²⁻, along with HA, substantially hindered MNZ decomposition. The trapping investigation confirmed that both radicals (SO₄˙⁻/˙OH) and non-radicals (¹O₂) were involved in the reduction of MNZ. The activation process provided additional confirmation of the redox cycling between Mn²⁺/Mn³⁺/Mn⁴⁺ during PMS activation, thereby elucidating its role in ROS generation and ultimate MNZ degradation. The Mn₃O₄/biochar catalyst exhibited stable and reliable catalytic performance along with low metal leaching across five cycles, thereby suggesting its suitability for repeated catalytic applications. Consequently, Mn₃O₄/biochar catalysts present a promising strategy for PMS activation in the degradation of antibiotics, thus offering an environmentally sound and dependable method for addressing environmental pollution.

Conflicts of interest

There are no conflicts to declare.

Data availability

Data will be available upon request.

Acknowledgements

The authors express their gratitude to the Princess Nourah bint Abdulrahman University Researchers Supporting Project (Grant No. PNURSP2026R12) and the Princess Nourah bint Abdulrahman University, Riyadh, Saudi Arabia. The authors extend their appreciation to the Deanship of Scientific Research, Islamic University of Madinah, Saudi Arabia, for funding this research work.

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Footnote

† equal contributions.

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