Activation of H₂O₂/PDS/PMS by Iron-Based Biochar Derived from Fenton Sludge for Oxidative Removal of 2,4-DCP and As(III)

Abstract

In this study, the catalytic performance of the Fenton sludge iron-based biochar catalyst (Fe@BC700), generated during the Fenton process, was investigated regarding its role in oxidizing 2,4-dichlorophenol (2,4-DCP) and As(III) from aqueous solutions in peroxymonosulfate (PMS), peroxydisulfate (PDS), and hydrogen peroxide (H₂O₂) systems. The characteristics of the as-prepared catalyst, operational parameters of H₂O₂/UV/Fe@BC700, PDS/UV/Fe@BC700, and PMS/UV/Fe@BC700 systems, and the kinetics of 2,4-DCP degradation were evaluated. Fe@BC700 exhibited excellent capabilities for activating persulfate and an outstanding oxidant performance as a heterogeneous photocatalyst under UV irradiation. Among the tested systems, PMS/UV/Fe@BC700 showed the highest oxidation capabilities for both 2,4-DCP and As(III) within 40 min. The total organic carbon (TOC) removal efficiency for 2,4-DCP was up to 95.9% in the PMS/UV/Fe@BC700 system. The presence of free radicals in the PMS/PDS system included ·OH, SO₄^·−, and ·O₂⁻, which were facilitated by both UV irradiation and the catalyst. The by-products generated during the PMS/UV/Fe@BC700 treatment were identified via LC-MS analysis, which showed that catalytic degradation substantially reduced the chronic and acute toxicity of 2,4-DCP intermediates. The present study demonstrates that the iron-based biochar derived from Fenton sludge exhibited remarkable persulfate activation capabilities and was highly effective in removing 2,4-DCP and As(III).

1. Introduction

The exponential increase in industrial production to meet daily human demands has resulted in the discharge of substantial amounts of effluent into water bodies worldwide. These discharges contain a diverse range of priority and emerging pollutants that primarily encompass aromatic organic compounds, their halogenated counterparts, and heavy metals such as arsenic (As), lead (Pb), cadmium (Cd), and chromium (Cr). Triclosan (TC), a common antimicrobial agent and 2,4-dichlorophenoxyacetic acid (2,4-D) herbicide, and its photodegradation by-products such as 2,4-dichlorophenol (2,4-DCP), have higher toxicity levels than triclosan itself [1,2,3,4,5]. Aqueous arsenic is present in most natural water as arsenate [As(V)] and arsenite [As(III)] [6,7,8]. As(III) is more toxic and mobile than As(V), with the toxicity of the former being approximately 26–60 times higher than that of the latter [9]. Compared to As(III), As(V) is more easily removed through adsorption or coagulation/precipitation [10]. Therefore, oxidizing arsenic from As(III) to As(V) is necessary for reducing its mobility and serves as a prerequisite process for several subsequent arsenic removal technologies [11]. The Fenton technology is widely recognized as the most efficient strategy for oxidizing As(III) to As(V) [9,12,13]. Under acidic conditions (pH 1–3), reactive oxygen species (ROS), such as ·OH and ${\mathrm{O}}_2\overline{·}$, present in the Fenton system effectively oxidize As(III) to As(V) [9,14]. However, under a neutral pH, the traditional Fenton process is highly inefficient due to iron speciation and its strong tendency to precipitate, resulting in a low utilization of H₂O₂. Peroxymonosulfate (PMS, HSO₅⁻) and peroxydisulfate (PDS, S₂O₈²⁻) are two stable and potent oxidants with redox potentials of 1.82 and 2.01 V, respectively [15,16]. Their activation during the degradation of organic compounds constitutes an important alternative technology [17]. However, due to the limited oxidation ability of PMS/PDS alone, direct oxidation of organic compounds has proven insufficient [18,19]. Therefore, various strategies such as using transition metals (e.g., Fe²⁺, Mn²⁺, Cu²⁺, and Co²⁺), thermal treatments, UV radiation, and carbonaceous-based catalysts can be employed to effectively activate PMS/PDS for ROS generation and accelerate pollutant degradation [15,20,21,22]. Heterogeneous catalysts containing transition metals exhibit pronounced electron transfer capabilities, efficiently activating PMS and PDS without the need for energy consumption [23,24]. Consequently, these catalysts offer substantial advantages in advanced oxidation processes (AOPs) [25]. The compatibility of iron-based catalysts with the oxidation of a wide range of pollutants and the facilitation of their separation after treatment has earned considerable attention [26,27,28]. However, practical applications of iron-based catalysts are restricted due to their limited pH range and slow conversion rate [29,30]. Biochar, with its abundant carbon content, rich surface functional groups, and high specific surface area, can serve as an activator for PMS and PDS and act as a carrier for iron-based catalysts [30,31,32,33].

Fenton sludge (FS) is an inevitable by-product generated during the Fenton treatment process, primarily consisting of ferric hydroxide, natural organic matter, solid impurities, and other substances [34,35,36,37]. Considering the imperatives of sustainable development and cost-effectiveness, FS has emerged as a promising resource for preparing heterogeneous iron-based biochar catalysts in water treatment with considerable practical application potential. Cho et al. [38] reported that the biochar produced from FS through pyrolysis at 900 °C exhibited an excellent adsorption capacity for cadmium, with a maximum adsorption capacity of 260.2 mg/g. Xia et al. [39] applied FS co-mingled with black liquor and under pyrolysis conditions produced a biochar catalyst for the catalytic degradation of rhodamine B through Fenton-like oxidation, which produced a nearly 100% removal rate within 10 min. A study conducted by Cui et al. utilized coal mine drainage sludge, which contains high concentrations of iron, to effectively remove Cd(II), Cu(II), Pb(II), and Zn(II) from mine water [40]. Therefore, iron-containing sludge can serve as a valuable resource for preparing high-performance catalysts, facilitating beneficial recycling practices, and mitigating the environmental impacts associated with its disposal. However, the degradation mechanism of Fe-based biochar has primarily been investigated for H₂O₂ processes [41,42,43], and limited research has been conducted on PMS-based/PDS-based AOPs for treating water contaminated with 2,4-DCP or oxidizing As(III) to As(V).

The objective of this study was to develop and assess an environmentally friendly Fenton-like reaction system. FS was selected as the catalyst source and synthesized Fe-based biochar (Fe@BC700) via the pyrolysis method; 2,4-DCP and As(III) were chosen as the target pollutants. The catalytic performance of the as-prepared catalyst was assessed under a series of conditions, including oxidant quantity, catalyst dosage, anion concentration, and pH. Furthermore, a comparative analysis of active species in H₂O₂, PMS, and PDS systems was conducted to elucidate the catalytic oxidation mechanism. Degradation intermediates during the treatment process were also analyzed along with the catalyst reusability and a toxicity assessment of degradation intermediates. In this study, FS was utilized to synthesize iron-based biochar, which serves as a heterogeneous catalyst in the persulfate photocatalytic system to facilitate the oxidation of 2,4-DCP and As(III), thereby offering a novel approach for resource utilization of FS.

2. Material and Methods

The raw FS for this study was obtained from Guitang Sugar Group Co., Ltd. (Guigang, China). To prepare the catalyst for the experiment, the FS samples were dried, ground, and subsequently sieved through a 100-mesh screen. The resultant FS was then stored in a sealed bag within a desiccator. Thereafter, 5 g of FS was introduced in a tube furnace, which was heated to the desired temperatures of 300, 500, and 700 °C under an N₂ atmosphere, with the final temperature being sustained for 90 min. The prepared catalysts were designated as Fe@BC300, Fe@BC500, and Fe@BC700.

In this study, 2,4-DCP and As (III) were used as simulated contaminants at 100 mg/L. A CEL-LB70 photoreactor (Zhongjiao Jinyuan, Beijing, China) with a UV cutoff filter (λ < 420 nm) was used to conduct simulated UV degradation experiments on these pollutants. In typical photocatalytic degradation experiments of simulated pollutants, the pollutant solution was mixed with a catalyst concentration of 0.5 g/L and adjusted to a pH of 7. Subsequently, the mixture was agitated in the dark for 60 min to ensure adsorption–desorption equilibrium. An oxidizer (H₂O₂, PMS, or PDS) was then added under illumination from a 500 W Hg lamp to start the catalytic reaction. Throughout the photocatalytic process, 1 mL aliquots of the reaction solution were collected at predetermined intervals and filtered through a 0.22 μm membrane, and 0.3 mL of methanol was subsequently added as a quenching agent to the filtrate. The residual concentrations of 2,4-DCP or As(III) were measured using HPLC-QTOF (Thermo Fisher, MA, USA) and an atomic fluorescence spectrophotometer (Beijing Baode Instrument Co., Ltd., Beijing, China).

The aqueous stability of the catalyst is assessed by quantifying the iron concentration during the photocatalysis process. In the present study, 0.05 g catalyst sample was dispersed in 100 mL distilled water (pH 7). The mixture was subsequently agitated in the dark at 25 °C for 60 min, followed by 40 min of photocatalysis. The resulting supernatant was filtered using a 0.22 μm membrane. The concentration of the iron ions in the filtrate was determined via atomic absorption spectrometry (ContrAA800D, Thuringia, Germany). Post-analysis, the residual powder was subjected to centrifugation, reactivated at 700 °C under a nitrogen atmosphere, and reused for subsequent experiments as well as XRD and XPS characterization. All tests were concurrently performed in triplicate.

The specifics concerning the chemical reagents, catalyst characterization methods, testing procedures, and techniques for trapping active species employed in this study are provided in the Supporting Information Text S1–S3.

3. Results and Discussion

3.1. Characterizations and Chemical Properties of Catalysts

The XRD patterns of samples obtained at various pyrolysis temperatures were analyzed to investigate the influence of temperature on the structural characteristics of catalysts (Figure 1a). The raw FS displayed no characteristic peaks associated with the iron-containing phases. Fe@BC300 exhibited a distinct peak at 2θ of 26°, which was attributed to the d(012) of α-Fe₂O₃ (JCPDS 33-0664) [44,45]. For Fe@BC500, new diffraction peaks appeared at 2θ = 35.5°, and 62.5°, corresponding to the d(311) and d(440) of Fe₃O₄ [46]. The characteristic peaks of α-Fe₂O₃ showed a significant increase after reaching a pyrolysis temperature of 500 °C. Fe@BC700 showed new typical peaks at 43°, corresponding to the d(220) of Fe₃C (JCPDS 65-2411), and at 44.6° and 65°, corresponding to the d(110) and d(220) of Fe⁰ (PDF87-0721) [39,47]. Fe₃O₄ constituted the main crystalline phase on the surface of Fe@BC500, whereas Fe@BC700 consisted of a mixture of Fe₃O₄ and Fe⁰.

The impacts of pyrolysis temperatures on the functional group types on biochar were revealed by the FTIR spectrum (Figure 1b). The intensity of the broadband observed at 3500 cm⁻¹ decreased significantly with increasing pyrolysis temperature. This broadband was associated with the –OH vibration in the inorganic compound, indicating the decomposition of water and carboxylic in sludge during the carbonization process [48,49]. Subsequently, emerging bands at 543 cm⁻¹ and less than 700 cm⁻¹, assigned to Fe-O and Fe₃O₄, respectively, were observed after the catalyst formation exceeded a temperature of 500 °C [50,51]. This phenomenon further suggested that the functional groups of Fe species in both Fe@BC500 and Fe@BC700 were influenced by the high pyrolysis temperatures resulting from carbon thermal reduction processes.

The surface compositions of Fe@BC300, Fe@BC500, and Fe@BC700 were next analyzed via XPS (Figure 1c,d). The O1s of the catalyst (Figure 1c) was deconvoluted into three peaks: Fe-O (530–530.15 eV), C=O (531.39–531.50 eV), and C-O (533–533.12 eV) [41,51,52,53,54]. With an increase in pyrolysis temperature from 300 to 500 to 700 °C, the content of Fe-O increased from 13.87 to 21.18 to 24.71%, respectively (Table S2). The high-resolution Fe2p spectra (Figure 1d) reveal that Fe@ BC300 exhibits a peak at 710.56 eV, which is often associated with Fe2p of Fe³⁺. Fe@BC300 also shows a smaller peak at 719.29 eV, which is generally regarded as the satellite peak of Fe³⁺ [55]. Our results, combined with the previous XRD and FTIR spectra, showed that Fe@BC300 iron was mainly stabilized as trivalent iron compounds (e.g., Fe₂O₃, FeO(OH)) [56,57]. The high-resolution spectrum of Fe2p of Fe@BC500 had a peak of 709.19 eV, attributed to Fe²⁺ [58], thereby suggesting that some of the Fe³⁺ was reduced to Fe²⁺ when the pyrolysis temperature reached 500 °C [59]. In the high-resolution spectrum of Fe2p of Fe@BC700, two different peaks were observed at 706.22 and 719.02 eV, corresponding to Fe(2p 1/2) and Fe(2p 3/2) of Fe⁰, respectively [60]; this outcome is consistent with the XRD results. This finding indicates that when the pyrolysis temperature is further increased to 700 °C, Fe³⁺ or Fe²⁺ will be further reduced to Fe⁰. The Fe²⁺/Fe³⁺ ratios of the catalysts were estimated based on the relative peak areas of the characteristic peaks (Table S2). The slight decrease in Fe²⁺/Fe³⁺ to 0.26 after 90 min of carbonization at 700 °C compared to Fe@BC500 was attributed to the partial reduction of Fe²⁺ to the lower valence state Fe⁰ at 700 °C. The saturation magnetization rates of Fe@BC300, Fe@BC500, and Fe@BC700 were 2.52, 22.93, and 38.51 emu·g⁻¹ (Figure S2), respectively, which is consistent with the analysis results for the XPS spectrum.

3.2. Degradation of 2,4-DCP Under Different Systems

The impact of the pyrolysis temperature on the catalytic efficiency and mineralization of the prepared samples was initially investigated. The adsorption of Fe@BC300, Fe@BC500, and Fe@BC700 was 1.66, 6.19, and 9.95%, respectively (Figure S2a), which are consistent with the specific surface area results of the materials (Figure S4,

Table 1. Specific surface area, average pore volume, and surface element content.

	Fe@BC300	Fe@BC500	Fe@BC700
Surface Area (m2/g)	5.27	21.75	122.47
Pore Size (nm)	0.95	1.09	1.30
Pore Volume (m3/g)	0.0012	0.0059	0.049

). After 40 min of UV irradiation, Fe@BC700 exhibited the highest 2,4-DCP removal rate (94.93%) compared to Fe@BC300 (53.06%) and Fe@BC500 (55.23%) under the same reaction conditions. Fe@BC300 and Fe@BC500 exhibited TOC removal efficiencies of 25.44 and 44.54%, respectively. Notably, the highest TOC removal efficiency degradation was produced by Fe@BC700 (89.97%) (Figure S2b). Consequently, we selected Fe@BC700 for subsequent investigations.

The reaction parameters (oxidant concentration, catalyst dosage, initial pH, and inorganic ions) were investigated to further optimize the oxidation performance of Fe@BC700. The removal rates of Fe@BC700/PDS/UV and Fe@BC700/PMS/UV systems significantly increased with increasing pH (Figure 2a), reaching maximum degradation rates of 94.9 and 96.9%, respectively, at a pH of 7. The initial solution pH significantly influenced the conversion of SO₄^·− radicals to OH· in the PDS and PMS systems. Under acidic conditions, SO₄^·− was the predominant active species, whereas under neutral conditions, the reaction between SO₄^·− and OH⁻/H₂O generated OH·. Additionally, both OH⁻ and SO₄^·− synergistically contributed to contaminant degradation when the pH was close to 7 [15,61,62]. For the Fe@BC700/H₂O₂/UV system, the highest removal rate of 93.72% was produced at a pH of 3, followed by 83.84 and 82.87% at a pH of 5.0 and 7.0, respectively, which is consistent with other relevant research findings [63]. The effects of water matrix ions, such as Cl⁻, HCO₃⁻, H₂PO₄⁻, and NO₃⁻, on the removal rate of 2,4-DCP were also studied (Figure 2b). The 2,4-DCP degradation was significantly influenced by a concentration of 5 mM HCO₃⁻. This phenomenon occurred due to the consumption of ·OH and SO₄^·− radicals by HCO₃⁻; the chemical reactions that occurred in the H₂O₂, PDS, and PMS systems are represented by Equations (1) and (2) [64]. The order of the effect of anions on the degradation efficiency was HCO₃⁻ > H₂PO₄⁻ > Cl⁻ > NO₃⁻.

$${\mathrm{S}\mathrm{O}}_4^{·-}+{\mathrm{H}\mathrm{C}\mathrm{O}}_3^{-}\to {\mathrm{S}\mathrm{O}}_4^{2-}+{\mathrm{C}\mathrm{O}}_3^{·-}$$

(1)

$${\mathrm{H}\mathrm{O}}^{·}+{\mathrm{H}\mathrm{C}\mathrm{O}}_3^{-}\to {\mathrm{C}\mathrm{O}}_3^{·-}+{\mathrm{H}}_2$$

(2)

Figure 3a–c shows the effect of oxidizer concentrations (0.2 to 3.0 mM) on the degradation of 2,4-DCP and the corresponding apparent rate constant K values (Figure 3d). In the H₂O₂ system, the removal of 2,4-DCP increased with the H₂O₂ dosage (Figure 3a) as follows: 51.34% (0.2 mM), 64.41% (0.5 mM), 73.11% (1.0 mM), 82.87% (2.0 mM), and 88.75% (3.0 mM) within 40 min. Similarly, for the PDS and PMS systems, as the oxidizer concentration increased from 0.2 to 3.0 mM, the removal rate of 2,4-DCP increased from 57.79 to 95.23% and from 60.61 to 97.01%, respectively. However, there was no significant incremental pattern in either the removal efficiency and kinetic degradation rate of 2,4-DCP at C_PDS or C_PMS concentrations between 2.0 and 3.0 mM. This phenomenon can be attributed to the scavenging effect caused by an excessive concentration of oxidants in radical-based processes [65]. Sulfate and hydroxyl radicals were scavenged by an excessive PMS concentration, as demonstrated in Equations (3) and (4) [66]. In comparison to SO₄^·− and OH·, SO₅^·− exhibited a lower oxidizing power, resulting in a reduced system efficiency [67]. Thus, the optimum oxidizer concentration was 2.0 mM.

$${\mathrm{H}\mathrm{S}\mathrm{O}}_5^{-}+{\mathrm{H}\mathrm{O}}^{·}\to {\mathrm{S}\mathrm{O}}_5^{·-}+{\mathrm{H}}_2\mathrm{O}$$

(3)

$${\mathrm{H}\mathrm{S}\mathrm{O}}_5^{-}+{\mathrm{S}\mathrm{O}}_4^{·-}\to {\mathrm{S}\mathrm{O}}_5^{·-}+{\mathrm{H}\mathrm{S}\mathrm{O}}_4^{-}$$

(4)

The effect of Fe@BC700 usage on photocatalytic degradation activity was investigated (Figure 4a–c), and the corresponding apparent rate constant K is shown in Figure 4d. At an additional amount of 0.5 g/L, H₂O₂/UV/Fe@BC700, PDS/UV/Fe@BC700, and PMS/UV/Fe@BC700 exhibited degradation rate constants of 0.05, 0.16, and 0.52, respectively. With an increased addition amount of 0.6 g/L, the degradation rate constants were 0.05, 0.16, and 0.55, respectively. Notably, in the Fe@BC700/PMS/UV system, a full removal rate of 96.2% was produced within just 10 min. Therefore, the optimal experimental conditions were as follows: the usage of Fe@BC700 was 0.5 g/L, the oxidizer concentration was 2 mM, and there was no need to adjust the pH value (pH = 7).

To compare the photocatalytic oxidation properties of Fe@BC700 for organic pollutant degradation, different catalyst types were employed in the dark and under UV irradiation at a pH of 7 (Figure 5). When H₂O₂, PDS, and PMS were used individually, only 7.25 (Figure 5a), 10.91 (Figure 5b), and 10.40% (Figure 5c) of 2,4-DCP, respectively, were removed after 40 min. However, under identical experimental conditions, UV irradiation was applied for 40 min resulting in the removal of 42.62 and 44.63% of 2,4-DCP through PMS and PDS activation, respectively; thereby, surpassing the degradation produced solely by H₂O₂ activation. In contrast, when Fe@BC700 was simultaneously used with H₂O₂, PDS, and PMS, the removal rates were 48.14, 60.96, and 64.12%, respectively, which exceeded those obtained from oxidants/UV systems. This phenomenon can be attributed to the high specific surface area of the support material, biochar (Table 1), which enables effective adsorption of 2,4-DCP. This adsorption increases the likelihood of interactions between 2,4-DCP and reactive oxygen species (ROS) generated by Fe@BC700 and oxidants [68]. The highest degradation efficiency (92.9%) was demonstrated by PMS/Fe@BC70 after 5 min of UV irradiation, and complete degradation was observed after 40 min. This degradation efficiency was nearly 1.59 times higher than the degradation efficiency of the PDS/UV/Fe@BC700 system, and 3.16 times higher than that of the H₂O₂/UV/Fe@BC700 system. The oxidation rate constants (k) (Figure 5d) decreased in the following order: PMS/UV/Fe@BC700 (k = 0.52 min⁻¹) > PDS/UV/Fe@BC700 (k = 0.16 min⁻¹) > H₂O₂/UV/Fe@BC700 (k = 0.04 min⁻¹). The oxidation efficiency was further compared in different reaction systems using total organic carbon (TOC) measurements (Figure 5e). As anticipated, the TOC removal rates were 95.9 and 90.40% for the PMS/UV/Fe@BC700 and PDS/UV/Fe@BC700 systems, respectively, which were nearly 4.31 and 4.06 times higher, respectively, than that of the H₂O₂/UV/Fe@BC700 system.

In addition, the ecotoxicity of 2,4-dichloropropanol and its degradation intermediates in the PMS/UV/Fe@BC700 system is shown in Figures S5 and S6 and Table S3. The chronic toxicity assessment revealed that 2,4-DCP displayed toxicity towards Daphnia, algae, and fish; however, following a specific duration of degradation in the PMS/UV/Fe@BC700 system, most intermediates significantly diminished in toxicity. The values of LC50, EC50, and ChV all exhibited an increase, indicating a substantial reduction in the toxicity of 2,4-DCP following the treatment with the PMS/UV/Fe@BC700 system.

The photocatalytic oxidation efficiencies of As(III) by Fe@BC700 under different oxidation systems are shown in Figure 6. In the presence of H₂O₂, PDS, and PMS alone, a minimal oxidation of As(III) was observed within 40 min. However, after UV irradiation for 40 min, the oxidation rates reached 20.49, 30.52, and 34.31% for H₂O₂, PDS, and PMS, respectively. When both Fe@BC700 and oxidants were present in the As(III) solution, the dark absorption of As(III) on Fe@BC700 reached 18.50% within 60 min, and the adsorption capacity of Fe@BC700 was 0.025 mmol As(III)/g. The pyrolysis temperature positively influenced the specific surface area of the catalyst (Table 1). Fe@BC700 exhibited the highest value of 122.5m²/g, indicating its potential for the partial removal of As(III) through physical absorption. Although the adsorption capacity of Fe@BC700 was lower than that of other biochar-based materials, such as ferrihydrite-modified biochar (1.301 mmol As(III)/g), zero-valent modified biochar (0.05 mmol As(III)/g), and Fe-Mn-La-impregnated biochar (0.35 mmol As(III)/g), the catalyst was nonetheless crucially influential in enhancing As(III) removal [11,69,70,71,72]. The adsorption of As(III) increases the likelihood of interactions between As(III) and ROS. However, the primary mechanism for As(III) removal by Fe@BC700 was catalytic oxidation rather than adsorption, as further demonstrated in the subsequent catalytic degradation experiments. Following 40 min of UV irradiation, the PMS and PDS activations oxidized 57.35 and 61.02% of As(III), respectively, surpassing the efficacy demonstrated with H₂O₂ activation. The highest oxidation rate (96.2%) was attained by the PMS/UV/Fe@BC700 system, followed by PDS/UV/Fe@BC700 (90.9%), H₂O₂/UV/Fe@BC700 yielding an oxidation rate of 79.9% at an initial reaction pH of 7. Almost all As(III) was oxidized within 5 min during the PMS/UV/Fe@BC700 process, exhibiting a catalytic efficiency similar to that observed in the degradation of 2,4-DCP. These results indicate the synergetic effects between Fe@BC700, UV light, and PMS, and Fe@BC700 served as a versatile catalyst for persulfate activation. Specifically, PMS exhibited a superior activation potential compared to PDS and H₂O₂; this result may have occurred due to the PMS’s asymmetric molecular structure, which enables easier activation than other oxidants such as H₂O₂ [15,16]. Moreover, the presence of Fe⁰ enhanced the catalytic oxidation of As(III) due to its higher activity. The combination of Fe⁰ and biochar synergistically affected the degradation of As(III) through PDS/PMS/H₂O₂ activation compared with the conventional adsorption method because the biochar provided a larger specific surface area and ensured a uniform distribution of active sites, which resulted in a higher degradation efficiency of As(III) [39,73,74].

3.3. Radical Quenching Experiments and EPR Studies

To further investigate the primary contributions of crucial ROS in the H₂O₂/UV/Fe@BC700, PDS/UV/Fe@BC700, and PMS/UV/Fe@BC700 systems, three radical quenchers were utilized to explore their effects. When tertiary butyl alcohol (TBA) and p-benzoquinone (p-BQ) were introduced into the H₂O₂/UV/Fe@BC700 system (Figure 7a) to probe for ·OH and ·O₂⁻, a significant decrease in the degradation rate of 2,4-DCP was observed, indicating that ·OH and ·O₂⁻ were the predominant reactive species in the H₂O₂/UV/Fe@BC700 system. However, the introduction of TBA, p-BQ, and ethanol (Eth) into the PMS/UV/Fe@BC700 system resulted in a slight decline in the inhibition of 2,4-DCP degradation. Notably, p-BQ exhibited a more pronounced inhibitory effect compared to TBA and Eth, indicating that the ·O₂⁻ radical had a stronger influence than the ·OH and SO₄^·− radicals in 2,4-DCP degradation. The additions of TBA, p-BQ, and Eth into the PDS/UV/Fe@BC700 system yielded similar results, as observed within the PMS/UV/Fe@BC700 system.

To further validate the presence of ROS in the three systems, electron paramagnetic resonance (EPR) spectroscopy was employed for ROS detection. After 10 min of UV irradiation, there was a significant increase in the intensities of four- and six-line peaks corresponding to •OH and SO₄^·− (Figure 8a). Importantly, both PDS/UV/Fe@BC700 and PMS/UV/Fe@BC700 systems exhibited stronger EPR peak intensities for SO₄^·− and •OH radicals compared to H₂O₂/UV/Fe@BC700 under the identical reaction time. The intensity of typical spectra for DMPO-·O₂⁻ EPR (six peaks, 1:1:1:1:1:1) is shown in Figure 8b, where the concentration of the ·O₂⁻ radical in PDS/UV/Fe@BC700 and PMS/UV/Fe@BC700 systems surpassed that in the H₂O₂/UV/Fe@BC700 system. The quenching experiment and EPR spectroscopy experiments confirmed that the •OH, SO₄^·−, and ·O₂⁻ radicals were the predominant ROS in the PDS/UV/Fe@BC700 and PMS/UV/Fe@BC700 systems. Moreover, the concentration of the SO₄^·− radical was lower than that of the •OH radical in all reaction systems primarily due to the reaction between SO₄^·− and H₂O, resulting in the generation of the •OH radical [75]. The PMS degradation process that showed the highest activity can be attributed to its having a higher oxidation potential and lower LUMO energy for a higher acceptance of electrons [76].

ROS formation in H₂O₂, PDS, and PMS was proposed based on these experimental results. Fe@BC700 is an iron-based biochar catalyst that contains ferric metastable metals, such as Fe⁰, α-Fe₂O₃, and Fe₃C. Fe²⁺ acts as an electron donor for H₂O₂, PDS, and PMS to generate free radicals. The generated Fe³⁺ can continue to react with H₂O₂, PDS, and PMS and reduce to Fe⁰/Fe²⁺ to complete the redox cycle (Equations (5)–(15)) [77,78,79,80,81].

$${\mathrm{F}\mathrm{e}}^0+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{2+}+2{\mathrm{O}\mathrm{H}}^{-}$$

(5)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{3+}+·\mathrm{O}\mathrm{H}+{\mathrm{O}\mathrm{H}}^{-}$$

(6)

$${\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{2+}+·{\mathrm{O}\mathrm{H}}_2+{\mathrm{H}}^{+}$$

(7)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{3+}+·{\mathrm{O}}_2^{-}+2{\mathrm{H}}^{+}$$

(8)

$${\mathrm{F}\mathrm{e}}^0+{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\to {\mathrm{F}\mathrm{e}}^{2+}+2{\mathrm{S}\mathrm{O}}_4^{2-}$$

(9)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\to {\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{S}\mathrm{O}}_4^{2-}+{\mathrm{S}\mathrm{O}}_4^{·-}$$

(10)

$${\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\to {\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{S}}_2{\mathrm{O}}_8^{·-}$$

(11)

$${\mathrm{F}\mathrm{e}}^{3+}+\mathrm{H}\mathrm{S}{\mathrm{O}}_5^{-}\to {\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{S}\mathrm{O}}_5^{·-}+{\mathrm{H}}^{+}$$

(12)

$${\mathrm{F}\mathrm{e}}^{2+}+\mathrm{H}\mathrm{S}{\mathrm{O}}_5^{-}\to {\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{S}\mathrm{O}}_4^{·-}+\mathrm{O}{\mathrm{H}}^{-}$$

(13)

$${\mathrm{F}\mathrm{e}}^0+2{\mathrm{F}\mathrm{e}}^{3+}\to 3{\mathrm{F}\mathrm{e}}^{2+}$$

(14)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{3+}+·{\mathrm{O}}_2^{-}$$

(15)

In Fe@BC700, the biochar component had multifaceted effects on the adsorption and oxidation of 2,4-DCP and As(III), complementing the catalytic oxidation performed by ferric metastable metals. Biochar contains active functional groups capable of adsorbing and immobilizing As(III) and 2,4-DCP. The FTIR spectrum (Figure S6) reveals changes in these functional groups after adsorption. Specifically, the transmittance of key peaks increased and shifted compared to unabsorbed Fe@BC700. Notable reductions were observed in the absorbance peaks at 1351 cm⁻¹, and 3400 cm⁻¹, indicating a decrease in the amount of hydroxyl groups on the Fe@BC700 surface [82]. Additionally, the carbonyl (C=O) stretching vibration band at 1600 cm⁻¹ highlights the presence of an effective adsorption site for 2,4-DCP and As(III) [83,84]. Subtle changes in the positions of the C-O-C and Fe-O peaks further suggest that these oxygenated functional groups interact with As(III) to form a new functional group, As-O at 946 cm⁻¹ [71].

$${\mathrm{B}\mathrm{C}}_{\mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}}-\mathrm{O}\mathrm{O}\mathrm{H}+{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\to {\mathrm{B}\mathrm{C}}_{\mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}}-{\mathrm{O}\mathrm{O}}^{·}+{\mathrm{S}\mathrm{O}}_4^{·-}+\mathrm{H}\mathrm{S}{\mathrm{O}}_4^{-}$$

(16)

$${\mathrm{B}\mathrm{C}}_{\mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}}-\mathrm{O}\mathrm{H}+{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\to {\mathrm{B}\mathrm{C}}_{\mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}}-{\mathrm{O}}^{·}+{\mathrm{S}\mathrm{O}}_4^{·-}+\mathrm{H}\mathrm{S}{\mathrm{O}}_4^{-}$$

(17)

$$\mathrm{A}\mathrm{s}\left(\mathrm{III}\right)+2{\mathrm{S}\mathrm{O}}_4^{·-}\to \mathrm{A}\mathrm{s}\left(\mathrm{V}\right)+2{\mathrm{S}\mathrm{O}}_4^{2-}$$

(18)

$$\mathrm{A}\mathrm{s}\left(\mathrm{III}\right)+2·\mathrm{O}\mathrm{H}\to \mathrm{A}\mathrm{s}\left(\mathrm{V}\right)+2{\mathrm{O}\mathrm{H}}^{-}$$

(19)

$$·\mathrm{O}\mathrm{H}/{\mathrm{S}\mathrm{O}}_4^{·-}/·{\mathrm{O}}_2^{-}+\mathrm{2,4}-\mathrm{D}\mathrm{C}\mathrm{P}\to \mathrm{d}\mathrm{e}\mathrm{g}\mathrm{r}\mathrm{a}\mathrm{d}\mathrm{a}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n}\mathrm{p}\mathrm{r}\mathrm{o}\mathrm{d}\mathrm{u}\mathrm{c}\mathrm{t}+{\mathrm{C}\mathrm{O}}_2+{\mathrm{H}}_2\mathrm{O}$$

(20)

Based on these findings, a possible mechanism for removing 2,4-DCP and As(III) using Fe@BC700-activated oxidant (H₂O₂, PDS, or PMS) was proposed. First, 2,4-DCP or As(III) is adsorbed on Fe@BC700. Subsequently, the ferric metastable metals on the Fe@BC700 react with H₂O₂ to generate ·OH or activate PMS or PDS to produce SO₄^·− radicals. Simultaneously, the biochar surface-COOH and biochar surface-OH (corresponding to C=O and −OH vibration bands at 1600 cm⁻¹ and 1351 cm⁻¹, respectively, see Figure 1b) act as electron transfer mediators, enhancing PMS/PDS activation to generate SO₄^·− (Equations (16)–(20)) [83,85]. These radicals, in turn, react with 2,4-DCP or As(III), leading to their degradation [86].

3.4. Stability Testing

The presence of oxidant and ultraviolet radiation during photocatalysis can enhance the leaching of metal ions by heterogeneous catalysts. The stability of the catalyst is a crucial factor that influences the degradation effect of the photocatalytic system. To assess the reusability of the Fe@BC700 catalyst, five consecutive recycling runs were performed under identical reaction conditions. The catalytic activities of Fe@BC700 in H₂O₂/UV/Fe@BC700, PDS/UV/Fe@BC700, and PMS/UV/Fe@BC700 systems after the five recycling instances are shown in Figure 9a. The catalytic activity of Fe@BC700 was maintained over three consecutive uses, with a slight reduction in the efficiency observed in the fourth and fifth uses. In the fifth use, Fe@BC700 exhibited removal efficiencies of 79.88, 89.32, and 89.9% in H₂O₂, PDS, and PMS systems, respectively. Figure 9b shows the leaching of iron in the three systems; PMS/UV/Fe@BC700 showed a lower iron leaching rate compared to H₂O₂/UV/Fe@BC700 and PDS/UV/Fe@BC700. Specifically, the amounts leached after the fifth use were 1.71, 0.54, and 0.39 mg/L in the H₂O₂/UV/Fe@BC700, PDS/UV/Fe@BC700, and PMS/UV/Fe@BC700 systems, respectively, corresponding to respective leached iron amounts of 3.42, 1.08, and 0.78 mg from 1 g of Fe@BC700. Figure S7 illustrates the contribution of dissolved Fe ions to the degradation of 2,4-DCP in the PMS, PDS, and H₂O₂ systems. After 60 min, the dissolved Fe ions accounted for 19.46–20.25% of the degradation efficiency. This result indicates that the ions’ role in the overall catalytic performance was minimal, with the heterogeneous catalyst Fe@BC700 serving as the primary contributor to catalytic activity. The stability of the Fe@BC700 was further confirmed by XRD (Figure 9c) and XPS (Figure 9d) analyses conducted on both fresh and fifth used samples. The XRD patterns of the fresh catalyst and the fifth reused catalysts exhibited no significant changes, except for a slight decrease in the intensity of characteristic peaks at 2θ = 35.5°, 44.6°, and 65.0°. Similarly, the XPS results demonstrated the stability of the catalyst, with the Fe²⁺/Fe³⁺ ratio remaining unchanged after the fifth application. After five degradation cycles, Fe@BC700 showed no obvious chemical changes and maintained a strong catalytic ability, which indicated its continued stability and effectiveness in the photocatalytic oxidation process [87,88].

Compared with other biochar types (as shown in

Table 2. Comparison of photocatalytic performance of Fe@BC700 with previously reported biochar types.

Photocatalyst	Photocatalytic Condition
	Light and Target Pollutant	Oxidation	Catalysts (g/L)	Time (min)	Iron Leaching (mg/L)	Removal %	Ref.
FVC	Vis, ofloxacin 10 mg/L	PMS	0.7	150	/	95.2	[89]
Hem/BC-5	Vis, TC 10 mg/L	H2O2	0.2	90	0.45	96.1	[90]
CFO1B3	Vis, RhB 50 mg/L	H2O2	1	20	1.02	100	[91]
WSB	Vis, CR 10 mg/L	H2O2	1	180	/	60	[92]
Fe@BC700	UV, 2,4-DCP 100 mg/L	PMS	0.5	40	0.78	100	This work
Fe@BC700	UV, As(III) 5 mg/L	PMS	0.5	40	0.78	96.2	This work

), Fe@BC700 prepared in this study exhibited excellent water stability and photocatalytic activity under UV light irradiation and activation by several oxidants.

4. Conclusions

This study focuses on the preparation of iron-based biochar catalysts from Fenton sludge using a simple one-step pyrolysis method for efficient degradation of 2,4-DCP and As(III). Fe@BC700 exhibited excellent persulfate activation capabilities and excellent oxidant performance as a heterogeneous photocatalyst under UV irradiation. The removal rates of 100.0 mg/L 2,4-DCP by H₂O₂/UV/Fe@BC700 and PDS/UV/Fe@BC700 were 82.87 and 94.93%, respectively, at pH = 7.0, an oxidant concentration of 2.0 mM, and a reaction time of 40 min, whereas PMS/UV/Fe@BC700 was completely degraded. Furthermore, a TOC removal rate of 95.9% was obtained. The oxidation efficiency for As(III) at a concentration of 5 mg/L was 96.2% under the same conditions. The Fe@BC700 catalyst exhibited excellent reusability and stability, allowing for five cycles of recycling. However, the presence of anions, particularly HCO₃⁻ and H₂PO₄⁻, had a detrimental impact on its performance. Scavenging experiments confirmed that the degradation of 2,4-DCP and As(III) in the PMS/UV/Fe@BC700 process was primarily driven by oxidative species, such as •OH, SO₄^·−, and ·O₂⁻ radicals. The synthesized iron-based biochar catalysts demonstrated a high potential in catalyzing the Fenton-like process for treating organic pollutants and As(III), making them highly promising materials for wastewater treatment and accomplishing waste management objectives.