Synthesis of Fe-Loaded Biochar Obtained from Rape Straw for Enhanced Degradation of Emerging Contaminant Antibiotic Metronidazole

Abstract

In this study, magnetic (Fe)-loaded biochar was successfully prepared by a simple impregnation pyrolysis method. Meanwhile, its degradation capability and mechanism for typical antibiotic metronidazole (MNZ) were systematically investigated under different conditions. The characterization of the synthesized material showed that the specific surface area, pore diameter, and pore volume changed significantly. Also, functional groups and metal element Fe were introduced on the surface of the biochar, leading to its better capability to activate peroxymonosulfate (PMS). The degradation experiments showed that the removal of MNZ in the Fe-BC/PMS system can reach up to 95.3% in 60 min under optimal conditions. Free-radical capture experiments showed that there were several active species of •OH, SO₄•⁻, •O₂⁻, and ¹O₂ present in the catalyst to synergistically degrade MNZ, among which SO₄•⁻ played a major role; it was also found that the material can be easily recycled and was still effective after several uses. Further, the main degradation pathways of MNZ include nitrohydroxylation, hydroxyethyl functional group deletion, carboxylation of the amino functional group of •OH, demethylation, oxidation, and carboxylation. It is obvious that the synthesized magnetic-loaded biochar, Fe-BC, generated from waste rape straw crops, shows high catalytic performance in pollutant degradation, providing an insight into the recycling potential of waste biomass in the catalytic field for pollutant removal.

1. Introduction

Since penicillin was introduced in 1942, hundreds of antibiotics have been synthesized to be widely used in disease control and animal husbandry and have become a significant source of environmental micropollution [1,2]. Globally, antibiotics are used in quantities as high as 100,000 to 200,000 tons [3] per year, leading to an increase in antibiotic residues in the environment [4,5]. Also, antibiotic resistance has become a concern that threatens public health and environmental ecosystem [6]. Metronidazole (MNZ), a widely used antibiotic, was reported to enter the environment through feces and sewage effluent [7,8]. Thus, control measures for emerging antibiotic pollution are required.

Current methods for treating antibiotic wastewater include biological, physical, chemical, and advanced oxidation processes (AOPs). AOPs are innovative and successful treatment [3,9] approaches that include ozone oxidation, UV/H₂O₂, the Fenton procedure, the photo-Fenton procedure, and others [10]. Antibiotic removal has also been accomplished using in situ chemical oxidation of persulfate [11]. It is possible to activate peroxymono sulfate (PMS) and peroxydisulfate (PDS) to form the free radical SO₄•⁻, which has a wide pH flexibility as compared to other oxidants [12]. Numerous techniques, including heat treatment, ultrasonication, ultraviolet light, radiation, alkaline addition, metal transition, and carbon material activation, can be used to activate persulfates. Because non-homogeneous catalytic activation is less expensive, reusable, recyclable, and efficient across a broad pH range, it is recommended over homogeneous catalysis.

Degradation of antibiotics by AOPs is a homogeneous or heterogeneous process that breaks down antibiotics with large molecular structures into small molecule products via means of highly oxidative reactive oxygen species (ROS) [13]. •OH- and SO₄•⁻-based AOPs are currently the most widely used method for the oxidative degradation of antibiotics [14]. It should be noted that SO₄•⁻-based advanced oxidation processes (SR-AOPs) are preferred for antibiotic management [15] over conventional •OH-based processes [16] due to their stronger oxidizing capacity and higher mass transfer efficiency [17,18].

These days, there is a lot of interest in the use of carbon-based catalyst materials to activate the PDS/PMS process for the treatment of antibiotics [19]. Synthetic carbon-based materials have achieved good results in activating PDS/PMS for the degradation of organic pollutants [20,21]. However, most of the carbon-based catalysts are poorly reusable. For example, Qin et al. [22] synthesized nitrogen-doped porous carbon for the activation of PMS to degrade 4-chlorophenol; 100% of 4-chlorophenol was removed on the initial use and decreased to 54% after five repetitions in a cycling experiment. According to Zhang et al. [23], following the fourth cycle experiment, the rate of bisphenol A (BPA) dropped by 30%. Although carbon-based catalyst materials have the disadvantage of low reusability, these disadvantages can be overcome through optimization. In order to regenerate flaws on the surface of carbon nanotubes, Cheng et al. [24] used sustained UV irradiation; after multiple cycles of irradiation, the majority of the carbon nanotubes’ active sites were repaired. The degradation effect of carbon-based compounds, which are commonly used in wastewater to treat antibiotics and other contaminants, is readily apparent;

Table 1. Application of carbon-based materials in degrading antibiotics.

Pollutants	Initial Pollutant Concentration	Biochar Dosage	PMS Concentration	Reaction Time	Removal Rate %	Bibliography
Tetracycline	20 mg/L	0.1 g/L	1 mM	2.5 h	96.5%	[25]
Sulfamethoxazole	5 mg/L	0.05 g/L	4 mM	1.5 h	100%	[26]
p-Chloroaniline	5 mg/L	3 g/L	2.5 mM	1 h	92%	[27]
Methyl para-hydroxybenzoate	15 mg/L	0.02 g/L	307 mg/L	30 min	100%	[28]
Bisphenol A	20 mg/L	0.4 g/L	0.4 g/L	20 min	100%	[29]

provides a few instances.

Magnetic biochar (MBC) is a composite biochar that has been given magnetic separation qualities by adding magnetic elements to it [30]. Conventional biochar has been successful in environmental remediation, but as mentioned above, in practice, recycling issues need to be considered [31]. MBC can solve this problem due to its magnetic separation properties [32,33,34]. In addition, photocatalysts can be applied to MBC in order to enhance its catalytic performance [35]. Currently, a wide variety of feedstocks are available for the preparation of MBCs, including waste biomass and iron-containing biomass wastes [36,37]. The process of creating magnetic biochar composites (MBCs) from biomass often involves introducing magnetic precursors to give the composites paramagnetic characteristics [38]. Also, iron-rich biomass wastes can be directly utilized for the synthesis of magnetic biochar without the need for additional magnetic precursors [39]. Azalok et al. [40] produced Mn-Fe co-doped palm seed mesoporous biochar for photocatalytic oxidative degradation of tetracycline (TC) by co-precipitation-calcination method. When UV and PMS were present together, 90% of TC was eliminated entirely within 120 min. The combined effect of Fe and Mn enhanced the production of ROS and increased the efficiency of degradation. With a high clearance rate of 96.5%, Luo et al. [41] successfully activated PMS to break down the antibiotic ciprofloxacin; after multiple cycles, the removal rate of 88.2% was maintained.

Therefore, in this study, a composite catalyst consisting of Fe-loaded rape straw biochar (Fe-BC) was created using impregnation pyrolysis. Meanwhile, the effectiveness of MNZ antibiotic breakdown was assessed. In detail, “The effects of catalyst addition, PMS concentration, anions interference, and different pH levels’ impact on MNZ breakdown in the reaction system was studied. Further, research was conducted to uncover the reaction mechanisms and identify the active components within the system”. Lastly, the degradation mechanisms of MNZ were examined using mass spectrometry (LC-MS). The reuse of waste biomass resources was achieved while degrading antibiotic pollutants in wastewater.

2. Materials and Methods

2.1. Materials

Rapeseed straw as biomass was obtained from rapeseed fields in a district of Jiangsu Province, China. Sodium pertechnetate (Guangzhou Haoxuan Chemical Co., Guangzhou, China). MNZ, isopropanol (IPA), p-benzoquinone (p-BQ), ammonium oxalate (AO), and L-histidine (L-His) (Aladdin Reagent Co., Shanghai, China). Acetonitrile, formic acid, methanol (MeOH), and ammonium nitrate (Wuxi Prospect Chemical Reagent Co., Wuxi, China). PMS, tert-butanol (TBA), Na₂S₂O₃•5H₂O, sodium hydroxide, sodium bicarbonate, sodium chloride, sodium dihydrogen phosphate, and anhydrous ethanol (Aladdin Reagent Co., Shanghai, China). Nitrogen (Wuxi Sheng ma Gas Co., Ltd., Wuxi, China). with 99.99% purity. All reagents and compounds used were of analytical reagent grade, and no purification process was carried out.

2.2. Preparation of Rape Straw Biomass

The biomass rape straw used in this investigation was obtained from a nearby farm. Firstly, the collected rape straw was repeatedly rinsed with deionized water, ventilated to dry, and subsequently dehydrated at 80 °C using a forced-air drying oven (GZX-9030 MBE, Bo Xun Industrial Co., Shanghai, China). The powder obtained was pulverized using a wall-breaker (YXA-4, China Jiangsu Shibo Co., Suzhou, China). The material was sieved using a 100-mesh sieve, then weighed and stored for future use.

2.3. Preparation of Sodium Ferrate Modified Biochar

Weigh 5 g of pretreated powdered biomass into a 250 mL wide-mouth conical flask. Add sodium ferrate solution (12 g/L) and stir the mixture on a magnetic stirrer (84–1A Beijing Haitian You cheng Technology Co., Beijing, China) at room temperature (set the speed of the magnetic rotor at 600 rpm) until the rape straw biomass and sodium ferrate solution are fully mixed. Following thorough mixing, the material was subjected to drying in a blast drying oven, resulting in the formation of a solid block. Afterward, the block was removed and crushed into a fine powder using a mortar and pestle. The mixed rape straw powder was then transferred to a crucible for volatile fractions and subjected to high-temperature pyrolysis in a tube furnace. The temperature was adjusted to 700 °C, increasing at a rate of 5 °C per min. The reaction was carried out for 2 h under nitrogen with limited oxygen. After the furnace temperature had been reduced to ambient temperature, the magnetic biochar was removed, sieved, ground, and thoroughly rinsed with deionized water and anhydrous ethanol to remove the ash produced during the pyrolysis process. The washed catalyst material was then dried in an oven and labeled as Fe-BC. Refer to Figure 1 for the preparation process, Blank rape straw biochar was prepared using the same procedure and labeled as BC.

2.4. Catalyst Characterization

The catalysts were analyzed using X-ray photoelectron spectroscopy (XPS) instrument (K-Alpha, Thermo USA, Inc., Waltham, MA, USA). This technique allowed for the examination of the elemental composition, chemical state, and energy band structure of the catalysts based on the positions and shapes of the peaks observed in the plots [42]. The experimental parameters included the use of AlKα rays as the excitation radiation source, a beam spot size of 400 μm, a binding energy adjustment using C1s (at 284.80 eV), an operating voltage of 12 kV, and a filament current of 6 mA. The surface morphology, particle size distribution, and particle composition of the composite were examined and analyzed using a scanning electron microscope (SEM) (Quanta FEG 250, FEI USA, Inc., Hillsboro, OR, USA). The molecular structure of the catalyst was investigated using Fourier-transform infrared spectroscopy (FT-IR) (Nicolet iS20, Thermo USA, Inc., Waltham, MA, USA) to identify the chemical bonds and functional groups present. The diffraction pattern obtained by X-ray diffraction (XRD) (SmartLab, Nihon Rikaku Corporation, Akishima-shi, Japan) was used to study the physical structure and material composition of the catalyst to see whether the elemental loading was successful or not (Cu target Kα rays, scanning speed of 2°/min, and specific scanning angle range of 5~90°). The catalysts were analyzed for their specific surface and porosity using the BET method. This involved measuring the N₂ adsorption/desorption isotherms using a adsorbometer (ASAP 2020 HD88, Micromeritics USA, Inc., Norcross, GA, USA).

2.5. Degradation Experiment

This study conducted all reactions using 250 mL conical flasks; 500 mg/L MNZ standard mixture was configured, 4 mL of the standard mixture was extracted, diluted, and fixed to 100 mL, and the conical flasks were gradually filled with the required amount of deionized water, as well as various kinds and amounts of biochar. The initial reaction solution had a pH of 7, the concentration of MNZ was 20 mg/L, the magnetic stirrer’s rotational speed was set to 600 rpm, and the reaction was conducted at ambient temperature (25 ± 3 °C). At certain intervals of time (0, 5, 10, 15, 20, 30, 40, and 60 min), a certain amount of specimen was extracted and filtered through 0.45 μm filter heads (Beekman Biological Co, Nanjing, China) and then put into centrifuge tubes with 0.5 mm of methanol quencher added. Afterward, ultraviolet spectrophotometry operating at 320 nm wavelength was used to determine the concentration of MNZ in the reaction solution.

2.6. Influencing Factors Experiment

To investigate the effect of different variables on MNZ degradation, different catalyst dosages (0.005 g, 0.01 g, 0.02 g, 0.03 g, and 0.04 g), PMS doses (0.5 mM, 1 mM, 1.5 mM, and 2 mM), pH levels (3, 5, 7, 9, and 11 adjusted by adding NaOH (1 M) and HNO₃ (1 M)), and anions (Cl⁻, NO₃⁻, HCO₃⁻, H₂PO₄⁻, and SO₄²⁻ at 1 mM, 5 mM, 10 mM, and 20 mM concentration levels) were considered. The degradation performance of Fe-BC on MNZ was evaluated by fitting the degradation kinetic model.

To find out if the Fe-BC is reusable, after the initial batch degradation experiments, the following cyclic experiments were set up. The recycled materials were collected and recovered after the reaction through magnetic separation, then washed three times with 10 mL of anhydrous ethanol and deionized water to remove any residual contaminants on the surface. Finally, the cleaned materials were transferred to a blast drying oven for drying. The substance was then employed once more to degrade MNZ, and the preceding procedures were carried out.

2.7. Free-Radical Scavenging Experiment

To determine the reactive agents involved in the degradation reaction system, 0.05 mol/L of TBA, MeOH, BQ, and L-His as the bursting agents of the active species were used to scavenge for the possible presence of sulfate free radicals (SO₄•^–), hydroxyl free radicals (•OH), superoxide radicals (•O₂^–), and singlet oxygen (¹O₂), respectively. In the degradation experiments, the bursting agent was added to the system following the addition of the catalyst and oxidizing agent. Finally, the concentrations of the target samples in the system were monitored after the addition of various bursting agents to analyze the significant role played by free radicals.

3. Results and Discussions

3.1. Morphology and Structure Analysis

The surface morphologies of BC and Fe-BC were observed by SEM analysis. As illustrated in Figure 2a,b, the surface of the BC presents a smooth appearance, with an overall honeycomb shape and a relatively neat arrangement of layers, and a certain number of pores existed on the surface of the material, which provided abundant surface sites for loading of the nano-metal particles. On the contrary, after modification with sodium ferrate, the morphology of the biochar became relatively disordered and fragmented, and the biochar that lost its porous structure appeared to be more enriched, as shown in Figure 2c. From Figure 2d, the surface of Fe-BC exhibited a significant presence of many irregular and scattered tiny nanoparticles, most likely Fe₃O₄ particles, and it was the loading of Fe₃O₄ particles that led to the activation of PMS in the modified catalyst in the reaction system. The SEM analysis revealed that the surface structure of BC underwent alterations, and iron oxide particles were effectively deposited onto the biochar surface.

3.2. Surface Structure Analysis

Figure 3 shows the pore size distribution curves (Figure 3b) and nitrogen adsorption and desorption curves (Figure 3a) for BC and Fe-BC. Tests for both adsorption and desorption were conducted in 77k liquid nitrogen. The samples belonged to type IV adsorption isotherms, according to the International Union of Pure and Applied Chemistry’s (IUPAC) classification of adsorption isotherms. The inflection point appeared in the low-pressure section P/P0 = 0.12, and the amount of adsorption in the low-pressure section rose slowly, indicating that there was little force between the sample’s surface and the nitrogen molecules and that the sample had a low microporous content. With the increase in relative pressure, nitrogen capillary condensation appeared in the pores of the sample, forming an H4-type hysteresis ring, and there was no saturated adsorption platform in the H4-type hysteresis ring, indicating that the pore structure of the test samples was irregular, mainly stacked slit pores.

The BC’s specific surface area is 355.009 m²g⁻¹, according to

Table 2. The Fe-BC and BC specific surface areas and associated pore size characteristics.

Samples	Specific Surface Area (m2g−1)	Average Pore Size (nm)	Micropore Volume (cm3g−1)	Mesopore Volume (cm3g−1)
BC	355.009	2.3042	0.1484	0.0820
Fe-BC	34.477	4.2783	0.0126	0.0243

. The large specific surface area of the BC provided enough adsorption sites for the loaded metal crystals.

According to the characterization results, it can be seen that after modification by sodium ferrate, the maximum nitrogen adsorption capacity of the sample located at P/P0 = 0.98 decreased significantly, and an 82% drop in total pore volume was observed. It is evident from the pore size distribution curve that the two samples’ pore sizes were dispersed throughout the 1–100 nm range, with the greatest several pore sizes being found at 3.82 nm. The pore volume of the micro- and mesopores of the samples was obviously reduced after modification, and it can be seen that the signals of the pore volume after modification were obviously reduced, and the specific surface area of the samples was drastically reduced. These phenomena can be attributed to the homogeneous dispersion of the materials during the pyrolysis process of sodium ferrate, which blocked some of the micropores on the surface of the BC, and therefore, a substantial amount of nanoparticles occupied the pore space of the biochar.

3.3. Materials Analysis

The components of the chemistry on the surface of Fe-BC were analyzed using X-ray electron spectroscopy, and the XRD spectra in Figure 4 show that the crystalline peaks with diffraction angles of approximately 26.6° and 59.8° for the unmodified biochar BC are attributed to the crystalline characteristic peaks (011) and (121) on the surface of the hexagonal quartz (SiO₂), and to the crystalline characteristic peaks (011) and (121) on the crystal surface of the hexagonal quartz (SiO₂). The characteristic peaks on the crystalline surfaces, referenced to Standard Drawing Card PDF#01-075-8320, are often considered to be the graphitized amorphous structure of biochar [43]. The crystallographic peaks with diffraction angles of approximately 29.4°, 39.4°, and 47.6° represent the (104), (102), and (018) crystal planes of the rhombic system calcite CaCO₃, respectively, referring to Standard Drawing Card PDF#01-075-8320. This is because a large amount of CaC₂O₄ •H₂O present in the rape straw will decompose at high temperatures to form CaCO₃. From the XRD characterization of the Fe-BC, the crystallographic peaks with diffraction angles (2θ) of the sample around 30.1°, 35.4°, 43.1°, 57.1°, and 62.7° are characteristic peaks. Characteristic peaks at 7° are attributed to the (220), (311), (400), (511), and (411) crystal planes of the cubic crystal system magnetite (Fe₃O₄). Crystallographic peaks at diffraction angles of approximately 30.1°, 19.8°, 32.8°, and 58.7° (referenced to standardized image card PDF#01-080-6402) and the characteristic peaks on the (111), (220), and (422) crystal planes of K₂Se (referenced to Standard Drawing Card PDF#04-003-6959) are attributed to the cubical crystal system, with diffraction angles of approximately 44.7°, 65.0°, and 82.3°, respectively. Referring to the Standard Drawing Card PDF#04-004-9051, the peaks of α-Fe appearing on the (110), (200), and (211) crystal planes are attributed to the cubic crystal system, and the XRD spectrum shows that the modified biochar samples contain the peaks of Fe₃O₄, K₂Se, and α-Fe, which suggests that the biochar is loaded with mainly Fe₃O₄ after being impregnated by the sodium ferrate solution in the pyrolysis process, proving that the composite has been successfully synthesized.

3.4. Functional Group Analysis

Infrared spectroscopy was carried out in order to analyze the graphitized structures of the BC and modified Fe-BC. The infrared absorption peaks between 1425 cm^–1 and 1545 cm⁻¹ in the spectra are distinct and recognizable, as shown in Figure 5. These peaks correspond to the carbon materials’ D-band (sp3-ordered carbon) and G-band (sp2-disordered carbon), which are the stretching and contracting vibrations of the C=O and C=C bonds, respectively. Among them, the C=C group enhances the graphitization of biochar and promotes electron transfer efficiency, and the presence of the C=O group enhances the self-decomposition of PMS, resulting in the production of ¹O₂ and •O₂⁻. The −OH stretching vibration is responsible for the broad band seen at 3420–3490 cm⁻¹, according to earlier research. The stretching vibration of C-O and C-O-C bonds is responsible for the infrared absorption peak that is located about 1070 cm⁻¹. The C-H stretching vibration is responsible for the usual peak at 2900 cm⁻¹ [44]. In addition, the broadband at 570 cm⁻¹ clearly corresponds to the asymmetric vibration of Fe-O bonds, which is similar to that of Fe₃O₄ nanoparticles [45]. This observation confirms the existence of Fe₃O₄ particles on the surface of Fe-BC, which aligns with the findings from X-ray diffraction research. Overall, the relatively high number of functional groups in modified Fe-BC is attributed to the strong oxidizing ability of sodium persulfate, which introduced other functional groups during the loading process.

3.5. Bonding State and Elements Analysis

X-ray photoelectron spectroscopy is employed to examine the material’s chemical composition. The whole X-ray photoelectron spectrum of Fe-BC is displayed in Figure 6a, where it is evident that the surface of the material is composed of elements like C, O, and Fe, as well as two clearly distinguishable peaks, C 1s and O 1s.

Figure 6b demonstrates the synthesis of C 1s into three subpeaks, which represent the carbon–carbon single bond (C-C 284.8 eV), carbon–oxygen single bond (C-O 286.25 eV), and carbon–oxygen double bond (C=O 289.20 eV), respectively [46]. Figure 6c displays the high-resolution spectra of O 1s. The peak at 530.3 eV corresponds to the iron–oxygen (Fe-O) bond, indicating that the O1s composition is in the O²⁻ state [47]. The Fe 2p peaks (2p_1/2 and 2p_2/3) in the samples are given in Figure 6d, and it can be seen that iron in Fe-BC exists in two oxidation states, where the Fe 2p_2/3 peak can be divided into two sub-peaks at 711.55 eV and 714.25 eV, and the Fe 2p_1/2 peaks can be divided into two sub-peaks at 725.25 eV and 729.05 eV, which corresponds to the two valence states of Fe²⁺ and Fe³⁺, respectively. The structure of Fe₃O₄ is Fe(III)[Fe(II)Fe(III)]O₄, which indicates that in Fe-BC, The XRD characterization results are consistent with the presence of Fe mostly existing as Fe₃O₄.

4. Catalytic Assessment

4.1. MNZ Degradation in Different Catalytic Systems

Owing to the biochar material’s porous structure, adsorption is inevitable during the degradation of MNZ, and the removal of MNZ is realized through the synergistic effect of adsorption and degradation. In this part of the experiment, different systems were explored (blank, BC, PMS, Fe-BC, and Fe-BC/PMS) to illustrate the degradation of MNZ in each of these systems.

From Figure 7, it can be seen that there was almost no degradation of MNZ in the blank sample, which indicated that MNZ was relatively stable in a natural environment and is impossible to accomplish self-degradation in a short time period, whereas the addition of BC (0.3 g/L) and PMS (1 mmol/L) alone had a certain effect on the removal of MNZ (20 mg/L), and in the reaction time of 60 min, the removal rates of MNZ were 9% and 14%, respectively, indicating that BC played a certain adsorption effect on MNZ in the system where only BC was added, but the adsorption ability of BC on MNZ was very limited. The degradation of MNZ by PMS through natural photocatalytic activation was also very weak in the absence of catalysts; however, after the Fe-BC was added to the system, the removal rate of MNZ was clearly improved, reaching over 20% in the reaction time of 20 min and then reaching 31.1% after 60 min. This indicates that the increased adsorption efficiency of the magnetically modified catalyst on MNZ was due to the abundance of active sites in the pores of the magnetic biochar surface, which enhanced the adsorption effect. With the simultaneous addition of Fe-BC and PMS in the system, the removal rate of MNZ reached 66.3% within 5 min and reached up to 95.3% in 60 min, implying that the addition of Fe-BC accelerates the reaction speed.

After analyzing each reaction system, it was discovered that the Fe-BC/PMS system degraded MNZ the most effectively, which was because the Fe²⁺ in the magnetic biochar activated the PMS, resulting in the peroxygen bond in peroxydisulfate O-O break, generating the sulfate radicals SO₄•⁻, it is capable of breaking the organic matter’s carbon chain to accomplish the goal of pollution degradation. As the reaction proceeds, the degradation rate of the reaction process gradually decreases, which may be due to the fact that Fe²⁺ is oxidized during the activation process of the PMS, generating precipitation of ferric hydroxide, as shown in Equation (1).

$$4\mathrm{Fe}(\mathrm{OH}{)}_2+{\mathrm{O}}_{2 }+2{\mathrm{H}}_2\mathrm{O}\to 4\mathrm{Fe}(\mathrm{OH}{)}_3$$

(1)

4.2. Catalyst Dosage’s Impact on MNZ Degradation

In order to examine the effects of different concentrations of Fe-BC catalyst on MNZ degradation, 100 mL of MNZ solution containing a concentration of 20 mg/L was dosed with 0.05 g/L, 0.1 g/L, 0.2 g/L, 0.3 g/L, and 0.4 g/L of catalyst, respectively. According to Figure 8, the efficiency of removing MNZ rose from 67% to 97.5% when the concentration of Fe-BC was raised from 0.05 to 0.4 g/L. The reaction rate increased from 0.03223 min⁻¹ to 0.64365 min⁻¹. When the catalyst dosage was in the lower range, there was insufficient catalyst to provide enough active sites; as a consequence, the PMS was unable to be adequately activated in order to produce free radicals.

4.3. Effect of PMS Concentration on MNZ Degradation

It was shown that the addition of Fe-BC alone had very little effect on the degradation of MNZ in the particular reaction system. Consequently, a major factor in the deterioration of MNZ was the concentration of PMS. While holding the other parameters constant, the effect of various PMS dosage concentrations on the MNZ degradation was investigated. Figure 9 demonstrates that varying PMS concentrations had a significant impact on MNZ degradation. The elimination of MNZ initially increased and then reduced as the PMS concentration increased. When the PMS concentration rose from 0.5 mM to 1 mM, the reaction rate constant increased from 0.06694 min⁻¹ to 0.39201 min⁻¹, and the removal rates of MNZ were 80.9% and 96%, respectively. The acceleration of MNZ degradation with increasing PMS concentration is due to the accelerated generation of SO₄•⁻ as the active species in the reaction system increases with rising PMS concentration. As can be seen in Figure 4, Figure 5, Figure 6, Figure 7, Figure 8 and Figure 9, when the PMS concentration was increased to 2 mM, the degradation rate in the first 15 min was improved compared to that in the case of low concentration, but from 15 min to the end of the experiment, the degradation efficiency of MNZ was 92.9%, which was lower than that in the case of PMS concentration of 1 mM, and the possible reason for this outcome is that when the concentration of PMS increases, the active sites on the material’s surface become saturated, hence restricting the formation of SO4•⁻; moreover, high concentration of PMS produces excessive SO₄•⁻, which can react with itself in quenching reaction and with S₂O₈²⁻, instead of being unfavorable to the degradation of MNZ, and the reaction is shown in Equations (2) and (3).

$${•\mathrm{SO}}_4^{-}+{•\mathrm{SO}}_4^{-}\to {{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}$$

(2)

$${•\mathrm{SO}}_4^{-}+{{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}\to {\mathrm{SO}}_4^{2-}+{•\mathrm{S}}_2{\mathrm{O}}_8^{-}$$

(3)

4.4. Effect of pH on MNZ Degradation

Variations in pH will impact the production of unpaired electrons in the reaction system, thereby influencing the catalytic breakdown of MNZ. In this current investigation, the initial pH of the MNZ solution was adjusted to 3, 5, 7, 9, and 11 by adding a certain amount of 1 mM HNO₃ and 1 mM NaOH solution, and the initial pH value of the MNZ pollutant was determined to be 6.9. The findings indicated that when the pH value climbed from 3 to 11, the efficacy of MNZ removal reduced from 97.5% to 85.4%, and the reaction rate decreased from 0.61021 min⁻¹ to 0.09071 min⁻¹, respectively. Figure 10 clearly demonstrates that the degrading efficiency of the process was higher under acidic circumstances compared to alkaline conditions, which was attributed to the presence of excessive hydroxide ions in the pollutant solution under alkaline conditions, which covered the active sites on the catalyst surface, generated a large number of hydroxides and inhibited the generation of free radicals, thus reducing the catalytic activity of the modified biochar, whereas metal ions in the magnetic biochar were guided to migrate to the liquid phase under acid conditions, activating the PMS to generate more SO₄•⁻, which was more favorable for the degradation of MNZ. Under the acidic condition, the metal ions in the magnetic biochar were guided to migrate to the liquid phase, activating the PMS to generate more SO₄•⁻, which was more favorable to the degradation of MNZ [48]. The aforementioned results indicated that the Fe-BC/PMS system successfully degraded MNZ throughout a wide range of pH values, with better performance observed under acidic conditions in comparison to alkaline conditions.

4.5. Effects of Inorganic Anions on MNZ Degradation

The presence of inorganic anions including nitrate (NO₃⁻), bicarbonate (HCO₃⁻), dihydrogen phosphate (H₂PO₄⁻), sulfate (SO₄²⁻), and chloride (Cl⁻) will compete with MNZ for the adsorption sites on the surface of the magnetic biochar, on the other hand, these anions will scavenge the active sites, and the addition of some anions may also cause an increase in the pH value in the system, leading to the precipitation of the generated hydroxide on the composite surface and thus cover the active sites, resulting in a decrease in the activation effect. Figure 11 demonstrates that the presence of anions hindered the efficiency of catalytic degradation, with the inhibitory effects following a specific order: HCO₃⁻ > SO₄²⁻ > HPO₄²⁻ > Cl⁻ > NO₃⁻. Among these, the addition of HCO₃⁻ had an obvious inhibitory effect; the degradation of MNZ was reduced by 17.6% in 60 min with 20 mmol/L HCO₃⁻. From Equation (4), it can be seen that HCO₃⁻ will inhibit the generation of sulfate radicals during the degradation process, producing the weaker anionic radical HCO₃⁻ and thus lowering the catalytic efficiency, and from Equation (5), HCO₃⁻ can also react with S₂O₈²⁻ to reduce the catalytic efficiency [49,50]. In addition to the poor stability of HCO₃⁻, the low reactivity of MNZ on it also leads to low removal efficiency. As shown in Equations (6)–(8), the inclusion of Cl⁻, NO₃⁻, and HPO₄²⁻ ions had a minor inhibitory impact on the oxidative degradation of MNZ. This effect can be due to the interaction between the inorganic anions and the reactive radicals necessary for degradation, resulting in the formation of less reactive species.

$$•\mathrm{S}{\mathrm{O}}_4^{-}+\mathrm{HC}{\mathrm{O}}_3^{-}\to \mathrm{S}{\mathrm{O}}_4^{2-}+•\mathrm{HC}{\mathrm{O}}_3$$

(4)

$${\mathrm{S}}_2{\mathrm{O}}_8^{2-}+\mathrm{HC}{\mathrm{O}}_3^{-}\to \mathrm{HC}{\mathrm{O}}_4^{-}+\mathrm{S}{\mathrm{O}}_4^{-}+\mathrm{S}{\mathrm{O}}_3^{-}$$

(5)

$$\mathrm{C}{\mathrm{l}}^{-}+•\mathrm{S}{\mathrm{O}}_4^{-}\to •\mathrm{Cl}+\mathrm{S}{\mathrm{O}}_4^{2-}$$

(6)

$$\mathrm{N}{\mathrm{O}}_3^{-}+•\mathrm{S}{\mathrm{O}}_4^{-}\to •\mathrm{N}{\mathrm{O}}_3+\mathrm{S}{\mathrm{O}}_4^{2-}$$

(7)

HPO₄²⁻ + •OH → OH⁻ + •H₂PO₄⁻

(8)

4.6. Reaction Mechanism

It is generally believed that the active substances produced by catalyst-activated PMS are mainly SO₄•⁻. In order to confirm the speculation that Fe-BC activation of PMS produces SO₄•⁻, in this experiment, four substances, TBA, MeOH, p-BQ, and L-His, were added as the bursting agents for $•$OH, SO₄•⁻, $•$O₂⁻, and ¹O₂, respectively. Generally, MeOH is used as a bursting agent for oxidation reactions because it has high kinetic constants (k₁ = 9.6 × 10⁸ M⁻¹s⁻¹, k₂ = 2.4 × 10⁸ M⁻¹s⁻¹) and is widely used for scavenging $•$OH and SO₄•⁻ [51]. If SO₄•⁻ or $•$OH is present in the reaction system and plays a role in the oxidation reaction, the addition of methanol quenches both radicals and terminates the oxidation reaction. TBA is mainly used as a reactive quencher of $•$OH due to its high reactivity (k = 6.0 × 10⁸ M ⁻¹s⁻¹). p-BQ and L-His were used to test for the presence of single linear oxygen (¹O₂) and superoxide radicals ($•$O₂⁻), respectively [52]. Figure 12 displays the experimental results. Throughout the entire free-radical experiment, the catalyst dosage and PMS concentration were dosed at the ideal amount.

In the system without a free-radical bursting agent, the degradation efficiency of 20 mg/L MNZ by 0.3 g/L Fe-BC and 1 mmol PMS dosage could reach 95.3% in 60 min. The addition of TBA to the reaction system resulted in a drop in the degradation rate of MNZ to 85.1%, suggesting that the presence of •OH had a limited impact on the reaction system. After the addition of MeOH, the degradation effect was only 16.1%, which shows that SO₄•⁻ and $•$OH are both free radicals leading to the degradation of MNZ in the system, with SO₄•⁻ playing a major role in the degradation of MNZ. In addition, in the oxidation system with the addition of p-BQ and L-His, the degradation rate of MNZ decreased by 15.9% and 18.4%, respectively, which proved that ¹O₂ and $•$O₂⁻ also had a role in the removal of MNZ. Overall, in the Fe-BC/PMS reaction system, $•$OH, SO₄•⁻, $•$O₂⁻, and ¹O₂ were all involved, and the contributions were in the order of SO₄•⁻, ¹O₂, $•$O₂⁻, and $•$OH.

4.7. Cyclic Experiments

After the degradation experiments, the Fe-MBC composite was separated from the solution by applying a magnet field, washed several times with anhydrous ethanol and ultrapure water, dried in a blast drying oven, and then reintroduced into a new MNZ solution to repeat the experiment under the optimal reaction conditions (pH = 7, catalyst dosage of 0.3 g/L, PMS concentration of 1 mM, and MNZ concentration of 20 mg/L). After four consecutive repeated uses, the degradation efficiencies of the Fe-BC catalyst on MNZ were 95.3%, 91.9%, 86.8%, and 84.2%, respectively. The degradation efficiencies showed a small decreasing trend, and the reaction rate decreased from 0.33228 min⁻¹ to 0.07827 min⁻¹ with the repetition of the experiment. Overall, the material maintained a good degradation effect after four times of reuse. The decrease in the degradation efficiency of the catalyst after several cycles may be due to the fact that some intermediates generated during MNZ degradation adhered to the surface or pores of the catalyst, which masked the active sites, or it may be due to the fact that the magnetic stirring was always maintained during the experiments, which made the materials collide with each other, and ultimately resulted in the reduction in the metal leaching on the surface of the catalyst, thus leading to a decrease in the degradation effect. As depicted in Figure 13, the catalytic activity of the catalyst did not decrease significantly after four repetitions of the experiment, and the results implied that Fe-BC has good stability performance.

4.8. Degradation Pathways of MNZ

To obtain a deeper comprehension of the deterioration of MNZ in the Fe-BC/PMS system, the liquid chromatography–mass spectrometry (LC-MS) approach was applied to analyze the intermediate compounds and by-products of MNZ degradation, as depicted in Figure 14.

The Fe-BC/PMS system’s possible degradation pathways are depicted in Figure 15, with an emphasis on the key intermediates. In pathway 1, through direct attack of the C-N bond by free radicals and nitrohydroxylation, MNZ was converted to 1-(2-hydroxyethyl)-2-methyl-1H-imidazol-5-ol (P2, m/z = 142). Subsequently, P2 loses its hydroxyethyl functional group by reaction with $•$OH to yield 2-methyl-1H-imidazol-5-ol (P3, m/z = 98). Thereafter, P3 was converted to 2,5-dihydroxypyrrolidin-1-yl acetic acid (P4, m/z = 101) as a result of C-N bond breaking and carboxylation of the amino functional group of $•$OH [53]. Due to the unstable nature of the C-C and C=C bonds, they were oxidized by reactants and demethylated to form (E)-N-methylformamidine acid (P5, m/z = 59). In pathway 2, P2 is converted to 2-methyl-3-vinyl-3,5-dihydro-4H-imidazol-4-one by oxidative ring opening (P6, m/z = 124). Ethylglycine (P7, m/z = 103) was then generated via a carboxylation reaction [54]. Finally, P5 and P7 were oxidized to CO₂, H₂O and NO₃⁻ [55].

5. Conclusions

In this study, we successfully synthesized a magnetic biochar (Fe-BC) by impregnation pyrolysis and demonstrated that it can efficiently activate PMS to improve the degradation efficiency of MNZ. Characterizations of the Fe-BC composite evidenced that the main elements consisted of C, O, and Fe, with Fe mainly existing in the form of Fe₃O₄. The degradation process conformed to the proposed secondary reaction kinetics, and the degradation rate of MNZ could reach 95.3% in 60 min by Fe-BC/PMS under optimal conditions. The main active substances involved were SO₄•⁻, ¹O₂, $•$O₂⁻, and $•$OH, among which SO₄•⁻ played a dominant role in decomposing MNZ. In addition, the addition of different anions had a certain inhibitory impact on the catalytic degradation process. In conclusion, the iron-loaded magnetic biochar synthesized in this study could effectively activate PMS and showed good catalytic performance in degrading the antibiotic MNZ. The material has high reusability and stability. Resourceful utilization of waste biomass was achieved along with the effective removal of pollutants. This study provides a certain reference for biochar composite catalysts to synergies with PMS for the degradation of antibiotic pollutants and provides ideas for the design of new efficient and environmentally friendly composite catalysts.