Peroxymonosulfate Activation by Sludge-Derived Biochar via One-Step Pyrolysis: Pollutant Degradation Performance and Mechanism

Abstract

Municipal wastewater treatment relies primarily on biological methods, yet effective disposal of residual sludge remains a major challenge. Converting sludge into biochar via oxygen-limited pyrolysis presents a novel approach for waste resource recovery. This study prepared sludge-based biochar (SBC) through one-step pyrolysis of sewage sludge and applied it to activate peroxymonosulfate (PMS) for degrading diverse contaminants. Characterization (SEM, XPS, FTIR) revealed abundant pore structures and diverse surface functional groups on SBC. Using Acid Orange 7 (AO7) as the target pollutant, SBC effectively degraded AO7 across pH 3.0–9.0 and catalyst dosages (0.2–2.0 g·L⁻¹), achieving a maximum observed rate constant (k_obs) of 0.3108 min^–1. Salinity and common anions showed negligible inhibition on AO7 degradation. SBC maintained 95% degradation efficiency after four reuse cycles and effectively degraded sulfamethoxazole, sulfamethazine, and rhodamine B besides AO7. Mechanistic studies (chemical quenching and ESR) identified singlet oxygen (¹O₂) and superoxide radicals (O₂^•− ) as the dominant reactive oxygen species for AO7 degradation. XPS indicated a 39% reduction in surface carbonyl group content after cycling, contributing to activity decline. LC-MS identified five intermediates, suggesting a potential degradation pathway driven by SBC/PMS system. ECOSAR model predictions indicated significantly reduced biotoxicity of the degradation products compared to AO7. This work provides a strategy for preparing sludge-derived catalysts for PMS activation and pollutant degradation, enabling effective solid waste resource utilization.

1. Introduction

Activated sludge technology, a fundamental innovation in environmental engineering developed in the early 20th century, remains fundamental in modern wastewater treatment plants [1]. Rapid urbanization has significantly increased the generation of wastewater sludge, with data indicating that China’s sludge production has reached 3.86 million tonnes (dry solids basis) at 89% moisture content, projected to exceed 9 million tonnes by 2030 [2]. Sludge from municipal wastewater treatment plants is characterized by high moisture content, putrefaction potential, and odorous emissions, but also harbors substantial quantities of pathogens, parasite eggs, and recalcitrant heavy metals [1]. Improper disposal can lead to secondary environmental contamination, including soil and groundwater pollution. Traditional sludge management methods, including dewatering, landfilling, and incineration, pose significant environmental challenges, including land use conflicts, hazardous emissions, greenhouse gas generation, and eutrophication potential [3]. Valorizing biomass and sludge into clean energy represents a sustainable pathway for resource recovery [4]. Consequently, developing efficient and environmentally benign sludge treatment and resource recovery technologies holds significant practical importance. Recently, developing eco-friendly materials has emerged as a promising avenue for sludge valorization. Biochar, characterized by abundant feedstocks, low cost, high specific surface area, surface functional groups, and porosity [5], not only facilitates enhanced sludge valorization but also provides an effective catalytic platform, demonstrating significant application potential. The growing recognition of biochar’s value in sludge valorization supports the advancement of greener, more sustainable recycling pathways.

Advanced Oxidation Processes (AOPs), valued for their ability to mineralize pollutants into harmless compounds without secondary treatment, have emerged as highly promising methods for organic pollutant degradation [6]. AOPs generate highly reactive oxidants via mechanisms including photocatalysis, Fenton reactions, peroxymonosulfate (PMS) activation, ozonation, or their synergistic combinations. Compared to conventional Fenton systems, PMS/PS-based activation offers wider pH adaptability and higher redox potentials. These AOPs predominantly generate highly oxidizing radicals (hydroxyl [OH] and sulfate radicals [SO₄^•−]) or non-radical species (¹O₂) [7].

Radical generation in PMS/PS-based AOPs primarily occurs via homogeneous and heterogeneous activation. Homogeneous activation employs soluble metal ions but often causes secondary pollution and faces metal catalyst recovery challenges. Conversely, heterogeneous activation uses solid catalysts for oxidant activation, offering advantages like catalyst recyclability and efficient pollutant degradation, thus representing a primary research focus. Biochar, produced through thermochemical methods (pyrolysis, gasification, microwave-assisted treatment, hydrothermal carbonization), has attracted considerable attention [8]. Pyrolysis-derived biochar typically exhibits greater pore volume, higher surface area, and richer surface functional groups, making it the predominant production method [9,10]. During pyrolysis, biomass undergoes thermal decomposition under oxygen-limited conditions to form biochar. Key parameters, such as residence time, pyrolysis temperature, and heating rate, critically influence the physicochemical properties, structural characteristics, composition, and quality of the resulting biochar [11]. Biochar’s low cost, stability, non-toxicity, high surface area, porosity, and diverse active sites render it a promising catalyst for activating AOPs to remediate aqueous organic pollutants [12]. Recent applications demonstrate biochar’s capacity to activate oxidants (H₂O₂, PMS, PI, PDS, and ozone), generating reactive species for efficient pollutant degradation in AOPs [13]. Reactive species generation occurs through radical-mediated and non-radical pathways, facilitated by biochar’s unique structure and surface-loaded transition metals. In radical pathways, biochar activates peroxysulphate (PS) to generate SO₄^•− and ·OH, which exhibit high reactivity toward organic contaminants [14]. Non-radical pathways involve ¹O₂ generation and direct electron transfer process (ETP), offering selective oxidation mechanisms. Biochar’s structural defects (edge defects, vacancies) can interact with PS to form surface-bound intermediates that mediate electron transfer, promoting non-radical pathways [15]. As an electron mediator, biochar can directly transfer electrons from adsorbed pollutants to PS, enabling oxidation without radical generation [16]. Given the complex composition of activated sludge (microbial communities, residual organic matter, and inorganic components), the PS activation activity, selectivity, and reusability of SBC require systematic evaluation. The complex composition of SBC and its pollutant interactions necessitate investigation of its PMS activation performance and associated active species to optimize degradation efficiency.

In this study, wastewater treatment plant sludge was used to investigate the effects of different pyrolysis conditions (pyrolysis temperature, heating rate, residence time) on SBC performance in the PMS system. Comparative experiments were performed with N-doped, Fe-doped, and Fe-N co-doped biochar compared to undoped counterparts. AO7 served as a model pollutant to assess the PMS activation capacity of SBC. Environmental factors influencing AO7 degradation in the SBC/PMS system were investigated. SBC reusability and applicability to diverse pollutants were analyzed. AO7 degradation pathways mediated by SBC/PMS were elucidated. Degradation products were identified by LC-HRMS. Their toxicity was evaluated using ECOSAR predictive modeling and toxicological analysis.

2. Materials and Equipment

2.1. Chemicals and Reagents

All chemicals used in this study were purchased from the following suppliers: PMS, sodium persulfate (PDS), sodium periodate (PI), rhodamine B (RhB), sulfamethazine (SM2), sulfamethoxazole (SMZ), methanol (MeOH), tert-butanol (TBA), furfural (FFA), and disodium ethylenediaminetetraacetate (EDTA-2Na) were obtained from Aladdin Reagent Co., Ltd., Shanghai, China. Sodium thiosulfate pentahydrate (Na₂S₂O₃·5H₂O), urea (CH₄N₂O), 30% hydrogen peroxide, sodium carbonate (Na₂CO₃), sodium dihydrogen phosphate (NaH₂PO₄), and sodium perchlorate (NaClO₄) were purchased from Tianjin DaMao Chemical Reagents Factory, Tianjin, China. Benzoquinone (P-BQ) and sodium chloride (NaCl) were obtained from Sinopharm Chemical Reagents Co., Ltd., Shanghai, China. L-Tryptophan was purchased from Spectrum Chemicals, Inc., CA, USA. Humic acid(HA) and ferric chloride hexahydrate (FeCl₃·6H₂O) were obtained from Macklin Biochemical Technology Co., Ltd., Shanghai, China. L-Histidine (L-his) and AO7 were purchased from Shanghai Yuanye Biotechnology Co., Ltd., Shanghai, China. All reagents were of analytical grade and used directly without further purification. Ultrapure water (resistivity: 18.2 MΩ·cm), generated using a Milli-Q water purification system, was used throughout all experiments.

2.2. Preparation of SBC

In this study, the sludge was collected from the secondary sedimentation tank of a wastewater treatment plant in Panjin, China. To minimize the heterogeneity of the sludge, samples were taken bi-monthly and mixed for further treatment. Initially, the sludge was centrifuged at 10,000 rpm for 5 min, followed by drying at 105 °C for 2 h to remove moisture. The resulting solid samples were then ground into a powder (with the fresh sludge having a moisture content of approximately 72%) and sieved through a 100-mesh screen. Subsequently, the dried sludge was transferred to a tubular furnace (GSL-1100X, Hefei Kejing, China), where N₂ was continuously introduced to strictly limit the presence of oxygen during the pyrolysis process. The sludge was heated to the target temperature at a controlled rate under the N₂ atmosphere and maintained at the corresponding temperature for a specified duration to produce biochar material, achieving a yield of approximately 58%.

To investigate pyrolysis conditions’ effects on SBC, temperature, retention time, and heating rate were systematically varied. Using a fixed retention time (120 min) and heating rate (10 °C/min), three biochars were produced at different pyrolysis temperatures, designated as SBC-400, SBC-600, and SBC-800. With a pyrolysis temperature of 800 °C and retention time of 120 min, three additional samples were named according to their heating rates: SBC-2 °C/min, SBC-10 °C/min, and SBC-20 °C/min. When pyrolysis was conducted at 800 °C with a heating rate of 10°C/min, and retention times of 0, 30, 60, 120, and 180 min, the corresponding samples were labeled as SBC-0M, SBC-30M, SBC-60M, SBC-120M, and SBC-180M.

To investigate heteroatom doping effects, iron and nitrogen sources were, respectively, provided by FeCl₃·6H₂O and urea. These were mixed with sludge powder at 1:1 and 1:2 mass ratios, soaked for 12 h, and magnetically stirred. After oven-drying at 105 °C to constant mass, the mixtures were pyrolyzed under N₂ atmosphere (10 °C/min ramp to 800 °C, 2 h hold). The resulting doped biochars were designated as N1BC, N2BC, Fe1BC, and FeNBC according to dopant identity and concentration [17].

2.3. Characterization of SBC

Surface morphology and microstructure were characterized by scanning electron microscopy (SEM, Sigma 500, Zeiss, Oberkochen, Germany) equipped with energy-dispersive spectroscopy (EDS), and transmission electron microscopy (TEM, JEM-F200, JEOL, Tokyo, Japan). Elemental composition and distribution were analyzed via EDS mapping. Crystalline phases were identified by X-ray diffraction (XRD, 7000S, Shimadzu, Kyoto, Japan) using Cu Kα radiation (λ = 1.5406 Å). Scans were performed over a 2θ range of 5–90° with a step size of 0.02° at a scan rate of 6°/min. Surface chemical composition and bonding states were determined by X-ray photoelectron spectroscopy (XPS, ESCALAB 250Xi, Thermo Fisher, Waltham, MA, USA) with a monochromatic Al Kα X-ray source (1486.6 eV). Functional group modifications were characterized by Fourier-transform infrared spectroscopy (FTIR, IS5, Thermo Fisher, Waltham, MA, USA) with spectra recorded from 400 to 4000 cm⁻¹ at 4 cm⁻¹ resolution (32 scans). Structural defects and graphitization degree were assessed using Raman spectroscopy (inVia, Renishaw, New Mills, UK) with a 532 nm Nd: YAG laser source. Pore structure characteristics were determined from N₂ adsorption–desorption isotherms at 77 K (Micromeritics ASAP 2460, Norcross, GA, USA) using Brunauer–Emmett–Teller (BET, ASAP 2460, Micromeritics, Norcross, GA, USA) theory for surface area analysis. AO7 degradation intermediates were identified by LC-HRMS (Agilent 1290 UPLC/6550 QTOF, Agilent, Santa Clara, CA, USA).

2.4. Degradation of Pollutants by SBC/PMS

AO7, representing 15–20% of global azo dye consumption, was selected as the model pollutant due to its environmental persistence and documented ecotoxicity. Substrate selectivity was assessed by evaluating the activation efficiency of SBC/PMS toward SMZ, SM2, and RhB, representing pharmaceuticals and dyes with distinct molecular structures. The oxidant versatility of SBC was further examined using AO7 as the substrate in conjunction with PI, PDS, and H₂O₂ systems.

2.4.1. Degradation of AO7 by SBC/PMS

AO7 degradation experiments were performed in 150 mL amber conical flasks containing 50 mL of 20 mg/L AO7 solution. Predetermined quantities of SBC (0.4 g/L) and PMS (1 mM) were added sequentially. Reactions were initiated by placing flasks in a rotary shaker (25 °C, 150 rpm). Sodium thiosulfate (0.1 M Na₂S₂O₃) was prepared as quenching agent. Aliquots (1 mL) were withdrawn at 5, 10, 15, 20, 30, 45, 60, and 90 min intervals, and analyzed by UV-Vis spectrophotometry at 484 nm. Solution pH was adjusted (3.0–11.0) using 0.1 M H₂SO₄ or NaOH to evaluate pH effects on SBC/PMS performance. Effects of ionic strength, common anions, and HA were evaluated by adding respective reagents prior to reaction initiation. For regeneration studies, spent SBC was recovered by centrifugation, washed several times with ultrapure water and ethanol, and dried at 100 °C until constant mass for reuse. For potential applications in real water bodies, supernatant from sludge filtered through a circulating water vacuum pump, tap water, and river water were used in place of ultrapure water, with subsequent experimental steps consistent with those in Section 2.4.1. Radical quenching tests employed: TBA (100 mM) for ·OH/ SO₄^•−, MeOH (100 mM) for SO₄^•−, FFA(10 mM) for ¹O₂, L-his (10 mM) for ¹O₂, and p-BQ (1 mM) for O₂^•−. All experiments were conducted in triplicate with data reported as mean ± standard deviation (SD).

2.4.2. Oxidant Versatility and Substrate Selectivity of SBC/PMS Oxidant Selection and the Degradation of Other Pollutants by SBC/PMS

To evaluate the activation capability of SBC toward various oxidizers, H₂O₂, PDS, and PI were selected as alternative oxidants. The experiments commenced by adding SBC and oxidizers (at equivalent molar concentrations to PMS) into AO7 solutions, followed by incubation in a rotary shaker at 25 °C and 150 rpm. Subsequent procedures were consistent with those outlined in Section 2.4.1. To investigate the selective oxidation behavior of the SBC/PMS system, RhB, SM2, and SMZ were chosen as target pollutants. The experimental protocol for RhB degradation adhered to the methods described in Section 2.4.1. For the degradation of SM2 and SMZ, after dosing with SBC and oxidizers (equimolar to PMS), samples were collected at specific intervals (3, 6, 9, 12, 15, 20, 30, 45, and 60 min), quenched with 0.1 M Na₂S₂O₃, and quantified using HPLC with optimized detection wavelengths.

2.5. Analytical Methods

AO7 and RhB concentrations during degradation were quantified by UV-Vis spectrophotometry (UV-5200PC, Metash, Shanghai, China) at their respective λ_max values (484 nm for AO7, 554 nm for RhB). Solution pH was measured using a calibrated pH meter (FE28, Mettler Toledo, Zurich, Switzerland). SM2 and SMZ concentrations were determined by HPLC (UltiMate 3000, Thermo Fisher, MA, USA) equipped with C18 column (4.6 × 150 mm, 5 μm) using 0.1% formic acid/acetonitrile (60:40 v/v) mobile phase at a 1.0 mL/min flow rate. The degradation kinetics were analyzed based on a first-order kinetic model, with the apparent rate constant (k_obs) calculated according to the equation:

$$\ln \left({\displaystyle \frac{C_0}{C_t}}\right)=k_{obs}t$$

where C₀ (mg·L⁻¹) represents the initial concentration of the pollutant, C_t (mg·L⁻¹) is the concentration at time t, and k_obs (min⁻¹) is the pseudo-first-order rate constant.

The generation of reactive oxygen species (ROS) was monitored by electron spin resonance (ESR) spectroscopy (JES-FA200, JEOL, Tokyo, Japan) using 100 mM 5,5-dimethyl-1-pyrroline N-oxide (DMPO) and 2,2,6,6-tetramethylpiperidine (TEMP) as spin traps for O₂^•− and ¹O₂, respectively. Degradation products of AO7 were identified using LC-HRMS, and quantitative calibration curves were constructed with six-point standards, validated by triplicate measurements with R² > 0.995. Subsequently, toxicity assessments of AO7 and its degradation products were performed through ECOSAR modeling to evaluate potential ecological risks associated with the degradation pathway.

3. Results

3.1. Characterization Results of SBCs

The surface morphology and structure of SBCs at different pyrolysis temperatures were characterized using SEM (Figure 1a and Figure S1). SEM revealed that the pyrolyzed sludge gradually developed a layered structure, while SBC-800 exhibited a distinct multilayer structure with spherical particles on its surface. Furthermore, we conducted TEM on SBC at 800 degrees, and it corroborated these findings, showing a uniform distribution of C, N, and O within the SBC-800 particles (Figure 1b). Semiquantitative EDS showed the existence of various elements, with a higher proportion of Si, Al, and Fe among the inorganic elements (Figure S2). XRD analysis indicated that the primary crystalline phase in SBC-800 was quartz (SiO₂, PDF# 85-0798); no crystalline phases containing Fe or S were detected, suggesting that the surface spherical particles may be SiO₂ (Figure 1c).

Furthermore, FTIR shows that the content of functional groups varies with the increase in pyrolysis temperature. Specifically, the absorption bands corresponding to carboxyl groups (1626 cm⁻¹ and 1550 cm⁻¹) significantly decrease with the rise in target temperature [18]. This decrease indicates a gradual reduction in functional group content and enhanced aromaticity (Figure 1d). Raman spectroscopy revealed that all samples exhibited D-band (1330 cm⁻¹) and G-band (1560 cm⁻¹), with the I_D/I_G ratio increasing from 0.78 to 1.10, suggesting that pyrolysis promoted structural ordering of carbon materials (Figure 1e).

XPS deconvolution showed that the carbon content increased with temperature, while oxygen content fluctuated before decreasing, further indicating enhanced aromaticity due to pyrolysis (Figure 1f). The C1s spectrum peaks were assigned to C-C/C=C (283.92 eV), C-O/C-OH (284.67 eV), C=O (285.65 eV), and O=C-OH (288.05 eV). The peak observed at 284.6 eV was attributed to the presence of aliphatic or aromatic carbon species (Figure 2a). Additionally, the intensity enhancement [19,20] further corroborates that the aromaticity of biochar increases with rising pyrolysis temperature; carboxyl group content decreased with rising temperature, consistent with FTIR observations. N1s spectrum analysis revealed a significant presence of pyrrolic nitrogen only in SBC-400. At higher temperatures, the pyridinic nitrogen proportion decreased from 15.33% to 16.73%, while the graphitic nitrogen proportion increased from 44.29% to 55.78%, indicating that pyrolysis promotes nitrogen transformation and graphitization, enhancing aromaticity (Figure 2b). Collectively, these results demonstrate that pyrolysis improves the structural characteristics of SBCs while enhancing their aromaticity and degree of graphitization. The N₂ adsorption–desorption isotherms indicate the presence of both micropores and mesopores in all SBC samples. A substantial increase in N₂ uptake at low relative pressures (P/P₀) suggests a higher proportion of micropores in SBC-800 (Figure S3). Moreover, the specific surface area of SBC-800 is 151.67 m²/g, approximately 6.78 times greater than that of SBC-400 (Figure S4 and Table S1).

3.2. Activation Performance of Biochar Prepared Under Different Conditions

Study on Factors Influencing AO7 Degradation in the SBC/PMS System. In this study, AO7 served as a probe molecule to evaluate the activation capacity of various SBCs for PMS.

As shown in Figure 3a, under conditions where PMS and SBCs coexisted, SBC-400 and SBC-600 exhibited negligible AO7 removal, whereas SBC-800 demonstrated markedly enhanced degradation efficiency, reaching 96.5%. This observation aligns with previous findings by He et al. [21], indicating that the catalytic activity of biochar depends primarily on its surface active sites, which are strongly influenced by the pyrolysis temperature during preparation. To optimize SBC-800 preparation, effects of heating rate and retention time on the pseudo-first-order reaction rate constant k_obs for AO7 degradation were systematically investigated (Figure 3b). A heating rate of 10 °C min⁻¹ gave the highest k_obs (0.19187 min⁻¹). Furthermore, k_obs increased progressively with retention time, peaking at 120 min (0.19187 min⁻¹). Subsequently, we further investigated the influence of various heteroatom dopants on the activation of PMS by SBC in the degradation of AO7. Iron-modified biochar (FeBC) significantly inhibited the reaction; the k_obs for AO7 degradation with SBC-800 was approximately 12-fold higher than with FeBC (Figure 3c). This inhibition may be attributed to non-optimal iron doping, potentially causing active site occupancy or iron oxide aggregation on the biochar surface, reducing surface area and porosity [22]. Through experimental investigations, it was observed that Fe-N/BC exhibited a faster degradation rate compared to FeBC and NBC. This enhanced performance may be attributed to its higher hydrophilicity, which facilitates effective contact between the catalyst surface and the liquid phase, thereby inducing an elevated reaction rate [23]. Conversely, the doping of heteroatoms into biochar resulted in a suppressed degradation effect compared to undoped biochar. This inhibition may stem from inappropriate doping methods or excessively high doping temperatures, which hinder the effective incorporation of heteroatoms [24]. Considering both AO7 degradation efficiency and biochar production costs, SBC-800 was selected for subsequent experiments to investigate efficient AO7 degradation within the SBC/PMS system. Meanwhile, we compared the performance of the prepared SBC/PMS system in degrading AO7 with similar systems reported in other literature, as shown in Table S2. The comparison results indicate that our system’s degradation efficiency for AO7 is comparable to those reported by others, suggesting that the composition of different sludge samples and calcination methods may influence the system’s ability to activate PMS. Further investigation into the structure-activity relationships is necessary.

Figure 4a depicts the effects of SBC and PMS concentrations on AO7 degradation. As the SBC-800 dosage increased from 0.1 to 2.0 g·L⁻¹, k_obs increased ninefold (from 0.0342 to 0.3108 min⁻¹). Notably, complete AO7 decolorization occurred within 10 min at these dosages(0.2 to 2.0 g·L⁻¹). Furthermore, Figure 4b,c reveal that increasing PMS concentration from 0.25 to 2 mM caused k_obs to increase eightfold (from 0.0340 to 0.2922 min⁻¹). Conversely, increasing AO7 concentration from 10 to 40 mg·L⁻¹ progressively reduced k_obs by fivefold. Previous studies have reported that the dosage of SBC and PMS significantly influences pollutant degradation within SBC/PMS systems. The amount of SBC determines the availability of active sites, while increased PMS dosage enhances the generation of reactive species. Conversely, higher pollutant concentrations may lead to increased adsorption onto the catalyst surface, potentially hindering further degradation. Additionally, a greater pollutant load necessitates a corresponding increase in reactive species to achieve effective degradation, highlighting the importance of optimizing reagent dosages for maximal contaminant removal [25,26]. For subsequent experiments, SBC (0.4 g·L⁻¹), PMS (1 mM), and AO7 (20 mg·L⁻¹) were employed to examine key factors affecting AO7 degradation.

3.3. The Influence of Environmental Factors on the Degradation of AO7 in the SBC/PMS System

Figure 5a illustrates the strong influence of pH on PMS activation by SBC-800. Within pH 3.0–9.0, AO7 was decolorized within 10 min, with k_obs peaking at 0.2853 min⁻¹ (pH 3.0)—representing a sixfold increase over pH 11.0 (0.04357 min⁻¹). These results confirm that the degradation rate was highest under acidic conditions, identifying acidic conditions as optimal for PMS activation by SBC-800; this finding contradicts the results reported in [27], which indicate that the removal efficiency of AO7 increases with the initial pH value. Furthermore, in acidic media, the production of ¹O₂ in the system is five times lower than that in alkaline media, suggesting that the degradation of AO7 by the SBC/PMS system is predominantly non-radical in 27nature [28]. At pH 9 (near-neutral alkaline), degradation remained substantial (k_obs = 0.2110 min⁻¹), consistent with [29]. Ionic strength, common anions, and HA exhibited limited influence on AO7 decolorization (Figure 5b–d), with minimal changes in k_obs. At a HA concentration of 40 mg·L⁻¹, a significant inhibitory effect on the degradation of azo dye AO7 in the SBC/PMS system was observed. Research indicates that humic acid competes with target pollutants for ROS, thereby reducing degradation efficiency, with the inhibitory effect intensifying as HA concentration increases [30]. However, the concentration of humic acid in natural water bodies is typically much lower than the experimental conditions (e.g., surface water HA concentrations generally fall below 10 mg·L⁻¹). Consequently, this system may demonstrate favorable applicability in practical scenarios. Additionally, the system efficiently decolorized AO7 in tap and river water (Figure 5e), demonstrating practical applicability. On the contrary, the significant inhibition of AO7 degradation observed in the supernatant of sludge is primarily attributed to the presence of complex organic substances and a variety of anions and cations in the wastewater effluent. These constituents may hinder degradation performance by competing for reactive species or altering reaction pathways, thereby reducing the overall efficiency of AO7 removal.SBC-800 reusability was evaluated over four consecutive cycles. After four cycles, 95% AO7 could still be decolorized within 90 min (Figure 5f), confirming high reusability of the catalyst. Additionally, we conducted TOC measurements and observed that, during a reaction for 60 min, the TOC in the system did not decrease but rather increased. A possible reason is that the biochar catalyst, in the presence of PMS, can generate dissolved black carbon and soluble organic carbon (DOM). These DOM species can undergo self-polymerization on the radicals generated by PMS or interact with the degradation products of AO7, forming higher-molecular-weight species. These species directly contribute to TOC, thereby causing TOC to not accurately reflect the mineralization of AO7 itself [31].

3.4. Oxidant Versatility and Substrate Selectivity of SBC/PMS

In addition to evaluating SBC-800’s PMS activation capability, this study additionally assessed its performance with H₂O₂, PDS, and PI. Results reveal effective AO7 degradation via PDS activation by SBC-800 under neutral conditions, with k_obs following: PMS > PDS. PI activation was slower, yet complete decolorization was achieved within 45 min (Figure 6a,c). Conversely, H₂O₂ activation yielded minimal degradation, similar to H₂O₂ alone. This disparity may originate from: (1) differential reactive species generation (radicals vs. non-radicals); (2) varying site-specific oxidant activation capabilities; and (3) distinct active site involvement.

Given literature reports on PMS selectivity, we investigated substrate selectivity of SBC-800-activated PMS. Various organic compounds were selected as substrates for reactivity assessment. Results show RhB was decolorized within 25 min, while SMZ and SM2 were removed within 30 min (Figure 4b,d). This confirms marked substrate selectivity, with rapid RhB degradation and effective SMZ/SM2 removal. Previous studies have indicated that carbon-based materials predominantly operate via non-radical pathways, including singlet oxygen and ETP [32,33]. Moreover, non-radical mechanisms demonstrate notable selectivity and adaptability in complex aqueous environments [34], implying that the SBC/PMS system primarily facilitates pollutant degradation through non-radical pathways.

3.5. Study on the Degradation Mechanism of the SBC/PMS System

The formation and identification of reactive species in AOPs are crucial for elucidating the degradation mechanisms of organic contaminants. This study employed radical scavenging methods to precisely determine the predominant ROS in the SBC/PMS system and to investigate their roles. In the experiments, scavengers such as MeOH, TBA, p-BQ, FFA, L-his, and Trp were used to inhibit specific ROS, including ·OH, SO₄^•−, O₂^•−, and ¹O₂, see Figure 7a.

The results demonstrated that addition of these scavengers led to only approximately a 3% reduction in the observed k_obs for ·OH and SO₄^•−, indicating their limited contribution to the overall oxidation process [35]. Conversely, scavenging O₂^•− markedly suppressed the reaction, nearly eliminating O₂^•− -related pathways, which suggests that superoxide plays a critical role in the degradation mechanism [36]. Additionally, ¹O₂ is considered a key reactive species responsible for AO7 degradation. Prior research has identified O₂^•− as the primary ROS in biochar-activated PMS systems for dye degradation, with proposed pathways described by Equations (1)–(5). Therefore, the degradation mechanism in biochar-activated PMS predominantly involves non-radical pathways, although radical pathways also contribute to a certain extent, consistent with findings reported in references [37,38]. Equations (6) and (7) illustrate the pathways for ¹O₂ generation, providing a theoretical basis for understanding the underlying reaction mechanisms.

$$HS{O}_5^{-}\to S{O}_5^{2-}+H^{+}$$

(1)

$$S{O}_5^{2-}+H_{2}O\to S{O}_4^{2-}+H_2{O}_2$$

(2)

$$H_2{O}_2\to 2H{O}^{•}$$

(3)

$$H_2{O}_2+H{O}^{•}\to H{O}_2^{•}+H_{2}O$$

(4)

$$H{O}_2^{•}\to H^{+}+O_2^{•-}$$

(5)

$$O_2^{•-}+H{O}^{•}\to H{O}^{-}{+}^1{O}_2$$

(6)

$$O_2^{•-}+2H^{+}\to H_{2}O{+}^1{O}_2$$

(7)

Investigating pH effects on SBC/PMS revealed PMS deprotonation under alkaline conditions, generating hydroxyl radical precursor HO₂⁻. Subsequently, HO₂⁻ reacts to form superoxide radicals (O₂^•− Equations (8) and (9)). O₂^•− generation enables ¹O₂ formation at pH 9, consistent with mechanisms in Equations (6) and (7).

ETP constitutes another key mechanism for biochar-mediated oxidant activation and pollutant degradation(Figure 7b). Previous studies show redox potentials of graphene, biochar, and activated carbon are measurable electrochemically. EDTA-2Na served as a hole scavenger to probe ETP. AO7 degradation declined to 55.2% (5 mM EDTA-2Na) and 43.2% (10 mM) after 90 min with scavenger addition (Figure 5b), confirming the critical role of ETP. Previous report found high-temperature biochar enhances oxidative capacity via conductive carbon frameworks in non-radical pathways, thus the ETP plays a crucial role in this reaction system [39].

To further validate the conclusions derived from the chemical scavenging experiments, ESR spectroscopy was employed. DMPO and TEMP were used as specific spin traps to detect and identify ROS, including O₂^•− and ¹O₂ [40]. The ESR spectra (Figure 8a,b) revealed signals corresponding to these ROS, with signal intensities increasing significantly over the course of the reaction. These findings indicate that the generation of ¹O₂ and O₂^•− predominantly originates from the activation of PMS by SBC-800. However, conventional views often regard O₂^•− as an intermediate of ¹O₂, with minimal direct involvement in organic oxidation reactions [41]. This confirms that SBC-800 effectively facilitates PMS activation to produce these reactive species. Based on a review of relevant literature, it is evident that in the removal of pollutants using biochar-activated AOPS, the reactive oxygen species (ROS) generated during the reaction vary significantly depending on the origin of the biochar, the type of oxidant employed, and the nature of the pollutants. This underscores the critical importance of optimizing biochar activation conditions tailored to specific operational parameters. Further investigation into this area holds substantial potential for enhancing pollutant removal efficiency and advancing pollutant remediation technologies [41,42,43].

Figure 9a,b represent the original XPS spectra, which are plotted as stacked bar charts to facilitate comparison and analysis. Comparison of the C 1s XPS spectra before and after reaction (Figure 9d) reveals a significant reduction in oxygen-containing functional groups (C–O/C–OH and C=O) and a marked increase in the proportion of carboxyl groups (–COOH). These findings suggest that carbonyl content is associated with the activation performance of the SBC/PMS system, warranting further investigation. The decreased proportion of sp² C (aromatic/graphitic carbon) indicates increased oxidation and structural disorder, resulting in reduced surface area and diminished ETP capacity. Analysis of the O 1s spectra (Figure 9e) shows a considerable decrease in C=O groups, further confirming the critical role of carbonyl groups in the activation process. Consistent with the C 1s data, the ratio of O=C–OH increases significantly, implying that under strong oxidative conditions, SBC undergoes deep oxidation, primarily generating acidic groups such as carboxyls. This process can diminish the biochar’s ETP capacity and activation efficiency, and may involve structural modifications, including decarboxylation or deoxidation.

Figure 9f illustrates a decrease in pyridinic nitrogen proportion from 17% to 9%, a growth in pyrrolic nitrogen from 27% to 45%, and a concurrent decline in graphitic nitrogen content. This indicates that nitrogen-containing functional groups (pyridinic N and pyrrolic N) are progressively consumed during multiple reuse cycles due to oxidative degradation or leaching. Simultaneously, graphitic structures undergo thermal evolution or structural reorganization, resulting in a more ordered and stable configuration, characterized by reduced nitrogen content and a relative increase in graphitic carbon. These transformations compromise the catalytic activity and pollutant removal efficiency. Pyridinic N and pyrrolic N functionalities exhibit high reactivity, enabling ETP and activation of PMS to generate radicals, thereby facilitating catalyst-driven pollutant degradation. In contrast, graphitic nitrogen, with its higher electronegativity and smaller covalent radius, promotes ETP from adjacent carbon atoms, creating positively charged carbon species that interact with adsorbed PMS. The predominant degradation of AO7 via non-radical pathways suggests that graphitic nitrogen plays a crucial role in facilitating PMS activation [44,45].

Analysis indicates that multiple reuse cycles induce significant oxidative transformation of biochar surface chemical groups. Key redox-active sites—such as carbonyl groups, pyridinic, and pyrrolic nitrogen—are depleted and degraded during cycling. Simultaneously, extensive oxidation generates numerous inactive or low-activity acidic groups, notably carboxyl groups. The irreversible alteration of surface chemistry, combined with physical damage and leaching of inorganic components, is a major contributor to the observed decline in catalytic performance within the SBC/PMS system.

3.6. Intermediate Products and Toxicological Analysis of AO7 Degradation in the SBC/PMS System

Previous studies have demonstrated the rapid degradation capability of the SBC/PMS system for AO7. However, two critical issues remain pertinent in the context of practical water treatment: the identification of intermediate species generated during the reaction process and the assessment of the biotoxicity associated with these intermediates.

Degradation intermediates of AO7 were identified using HPLC-HRMS (Figures S5–S9), and based on these detected species, the proposed degradation pathway is outlined in Figure 10. Initially, the degradation of AO7 occurs primarily through the cleavage of the azo bond, resulting in the formation of intermediates M1 (sodium sulfanilamide) and M5 (1-amino-2-naphthol) [35]. M1 is subsequently oxidized to produce M2; however, the amino group in M2 is unstable and readily undergoes oxidation under aerobic aqueous conditions, leading to the formation of M3 and M4. Subsequently, M3 and M4 undergo desulfurization to yield M13. M5 is subjected to attack by ¹O₂, resulting in the formation of M6. The phenolic hydroxyl group in M6 is further oxidized to generate a quinone radical, leading to the formation of M7. M7 undergoes additional oxidation to produce M9, which is ultimately converted to M11. Meanwhile, M6 is also attacked by free radicals, subsequently oxidizing to form M8 and the monocyclic aromatic compound M10, which ultimately leads to the formation of M12. M12 is then hydrolyzed to yield M13, which is further transformed into M14. The subsequent cleavage of the aromatic ring results in the generation of carboxylic acids, alcohols, and alkanes. Ultimately, these degradation products are mineralized to CO₂ and H₂O [46].

Toxicity classification according to the Globally Harmonized System of Classification and Labelling of Chemicals: lg k ≤ 0 Very toxic; 0 < lg k ≤ 1 Toxic; 1 < lg k ≤ 2 Harmful; lg k > 2 Not harmful.

The acute and chronic toxicity of AO7 and its degradation intermediates were evaluated using the ECOSAR (Figure 11). Toxicity levels were classified according to the Globally Harmonized System of Classification and Labelling of Chemicals (GHS). AO7 was classified as harmful (Chronic Category 3) to green algae and daphnids. In contrast, most degradation products exhibited negligible acute toxicity to fish and were classified as non-toxic for chronic toxicity. As degradation progressed, LC50, EC50, and Chv values increased significantly for intermediates such as M5, M6, M7, M8, M9, and M10. Furthermore, green algae were identified as the most sensitive species to both AO7 and its degradation intermediates. Intermediate M11 was classified as toxic (acute) and very toxic (chronic). However, after further degradation to M14, the toxicity substantially decreased, with M14 classified as harmful (acute) and non-harmful (chronic). Although ECOSAR can provide an approximate range of toxicity, its prediction results still need to be confirmed by toxicity experiments.

4. Conclusions

This study developed a one-step pyrolysis method for the efficient synthesis of SBC and systematically investigated the influence of preparation conditions and heteroatom doping on SBC’s activation of PMS. Experimental results demonstrated that both calcination time and retention temperature significantly influence SBC performance, with high calcination temperatures proving critical. However, extended heating rates and retention times showed negligible impact on performance. Additionally, heteroatom doping significantly reduced PMS activation efficiency. Detailed investigation of AO7 degradation by the SBC/PMS system demonstrated sustained efficiency under diverse environmental conditions, highlighting its robust application potential. Preliminary evidence further suggests the system’s efficacy for degrading diverse pollutants. Mechanistic studies identified ¹O₂ and O₂^•− as the primary reactive species, directly correlated with SBC carbonyl group content. AO7 underwent degradation via radical and non-radical pathways, yielding transformation products with significantly reduced toxicity. Consequently, this study presents a readily synthesized SBC catalyst for effective PMS activation and pollutant degradation, embodying a viable waste-to-resource strategy.