Degradation of Acid Orange II by FeOCl/Biochar-Catalyzed Heterogeneous Fenton Oxidation

Abstract

In recent years, the rapid development of industry has led to the discharge of large quantities of pollutants, including harmful dyes, into water sources, thereby posing potential threats to human health and the environment. FeOCl and biochar have their own shortcomings as a mediator in the heterogeneous Fenton process. To make both materials useful, FeOCl supported on bamboo biochar (FeOCl/BC) was prepared by calcination using FeCl₃·6H₂O and bamboo powder as raw materials, and the composite’s catalytic activities were explored with acid orange II (AO-II) as the target pollutant. The degradation efficiency of FeOCl/BC composites on AO-II was determined by testing the mass ratio of FeOCl and BC, initial pH, temperature, H₂O₂ concentration, catalyst addition, addition of coexisting inorganic anions, and natural organic matter. The addition of biochar to FeOCl increased the activation of H₂O₂ to generate •OH for the removal of AO-II and accelerated the cycle of Fe³⁺/Fe²⁺. The removal rate of AO-II by the Fe₁C_0.2 composite was 97.1% when the mass ratio of FeOCl and BC was 1:0.2 (Fe₁C_0.2), which was higher than that of the pure components (FeOCl or BC) at pH = 6.1. Moreover, after five reuses, the Fe₁C_0.2 composite still showed high degradation activity for AO-II, with 83.3% degradation and low activity loss. The capture experiments on the active material showed that the removal of AO-II by the Fe₁C_0.2 composite was mainly dominated by •OH; however, •O₂⁻ and h⁺ played minor roles. The synthesized Fe₁C_0.2 composite could be applied for organic contaminants such as AO-II with high removal efficiency.

1. Introduction

The rapid expansion of the printing and dyeing industry has led to substantial discharges of dye-containing wastewater into aquatic ecosystems. Notably, these industrial effluents containing synthetic dyes pose significant environmental risks, demonstrating both ecotoxicological effects on aquatic organisms [1] and persistent threats to drinking water quality through bioaccumulation potential [2]. Industrial colorants are generally classified into two major categories based on their ionic characteristics: anionic acidic dyes and cationic alkaline dyes [3]. Among these, Acid Orange II (AO-II), a strongly acidic monoazo dye with high aqueous solubility, has been extensively employed as a coloring agent in leather processing and various industrial applications [4,5].

Many methods, such as adsorption [6,7,8], biodegradation [9,10,11], and chemical degradation [12,13,14,15,16,17], have been used to remove dyes from wastewater. Among these methods, Fenton technology, as a chemical treatment method, is regarded as superior because of its high degradation efficiency and strong mineralization ability. The traditional Fenton oxidation method, compared with other advanced oxidation processes, is favored because of its simple operation, rapid reaction, and ability to produce flocculation and other advantages [18]. Although Fenton oxidation technology has a better effect in treating organic pollutants in wastewater, there are slight shortcomings in practical applications; this is because the efficiency of Fe³⁺ returning to the Fe²⁺ reaction is not high, which limits the effective cycle of the Fenton reaction, reflecting the low utilization of H₂O₂ and incomplete degradation of organic pollutants [19,20].

Recently, Fenton technology-based iron oxychloride (FeOCl) has attracted much attention for its catalytic performance on the decomposition of organic contaminants and as a catalyst for the heterogeneous Fenton reaction [21,22,23]. However, its narrow band gap (~1.76 eV) leads to the fast recombination of photoexcited electrons and holes. This attribute limits photoelectrons, and their efficiency is low. In addition, FeOCl is effective in activating H₂O₂ only at an acidic pH [24]. One of the methods to overcome these limitations of FeOCl is to hybridize it with materials with super-catalytic capabilities, such as biochar (BC), which has been used extensively in Fenton reactions [25].

BC has a large number of pore structures, which can provide loading sites for photocatalysts and adsorb the target pollutants and bring the free radicals generated by photocatalysis into full contact with the target pollutants [26]. The electrical conductivity of BC can improve the separation and migration of electrons [27,28], and the combination of low-cost BC with iron-based photocatalysts can reduce the recombination of photogenerated e⁻ and h⁺. BCs are of great interest due to their good chemical stability, excellent electrical conductivity, and low cost [29]. Moreover, the electrons produced by the BC boost the surface iron redox in Fenton process [26]. Several studies have shown that biochar can enhance the biodegradation of antibiotics due to the enhancement of the contact between microorganisms and antibiotics via the adsorption onto biochar [30]. Inspired by the above results, we can speculate that after the introduction of biochar, the contact between free radicals and dye pollutants is enhanced. Although FeOCl for different reactions has been studied, studies on the catalytic activity of FeOCl/BC composites are very limited.

Compared to conventional Fenton catalysts (e.g., Fe₃O₄ or homogeneous Fe²⁺), the FeOCl/BC composite proposed offers two distinct advantages: (1) The FeOCl on BC enables electron-rich surfaces, significantly accelerating Fe³⁺/Fe²⁺ cycling. (2) The low-cost, biomass-derived BC framework enhances sustainability compared to synthetic supports. Here, we successfully incorporated FeOCl and BC in different ratios. The prepared FeOCl/BC composites were chosen to degrade AO-II by the Fenton reaction in the presence of H₂O₂. The effect of reaction parameters (temperature, initial pH, catalyst amount, and H₂O₂ consumption) and the addition of coexisting inorganic anions and natural organic matter on the efficiency of FeOCl/BC composites to degrade AO-II were investigated. Finally, the Fenton reaction mechanism of FeOCl/BC/H₂O₂ was proposed for the synergistic effect of FeOCl and BC.

2. Experimental Section

2.1. Materials

AO-II (≥99%), H₂O₂ (~28–30%), ferric chloride hexahydrate (≥99%), and anhydrous ethanol (ETOH) (≥99.8%) were purchased from Beijing Bailingway Technology Co., Ltd. (Beijing, China). Isopropanol (IPA), p-benzoquinone (p-BQ), and EDTA were obtained from Sinopharm Chemical Reagent Corporation (Shanghai, China). Other reagents were procured from Aladdin Chemical Reagent Corporation (Shanghai, China). Deionized water was used during the whole experiment.

2.2. Synthesis of Catalysts

2.2.1. Preparation of BC

The BC was synthesized following the published work of Li et al. [31] with minor modifications. The bamboo powder was repeatedly soaked in distilled water, fully washed and dried naturally, and then sieved through a 200-mesh sieve to obtain bamboo powder. The pretreated bamboo powder was placed in a quartz crucible and pyrolyzed under N₂ at 550 °C with a heating rate of 5 °C·min⁻¹ for 2 h to obtain the BC material.

2.2.2. Synthesis of FeOCl/BC Composites

The FeOCl/BC composite was synthesized as follows. In a typical fashion, FeCl₃·6H₂O and BC were weighed at mass ratios of 0.1, 0.2, 0.5, and 1.0, ground in a mortar until they were mixed well, transferred to a crucible, and then heated to 250 °C for 2.5 h. The black powder was washed three times with deionized (DI) water and absolute ethanol and then dried in a vacuum oven at 60 °C for 6 h. Last, the black FeOCl/BC composites were obtained. The corresponding photocatalysts were labeled as Fe₁C_X, where x is 0.1, 0.2, 0.5, and 1.0 for the mass ratios of FeCl₃·6H₂O and BC, respectively. The preparation process is schematically illustrated in Scheme 1. As a control, FeOCl was synthesized following an identical procedure to that of FeOCl/BC excluding the BC component [21].

2.3. Degradation Experiment Evaluation

Comprehensive details are provided in Text S1 of the Supporting Information.

2.4. Characterization

Comprehensive details are provided in Text S2 of the Supporting Information.

3. Results and Discussion

3.1. Degradation Efficiency of AO-II Under Different Reaction Systems

The removal rates (adsorption coupled with oxidation) of AO-II using FeOCl/H₂O₂, BC/H₂O₂, and Fe₁C_X/H₂O₂ systems were investigated (Figure 1a). Biochar (BC) demonstrates negligible adsorption capacity. FeOCl displays marked adsorption capabilities. The adsorption rate of the Fe₁C_X composites for AO-II slightly increased compared to the adsorption rate of the pure components. Consequently, the FeOCl/BC composite exhibits synergistic adsorption properties through interfacial charge transfer mechanisms, where the FeOCl component provides active adsorption sites while BC enhances surface accessibility. The removal efficiencies of the catalysts for AO-II in the presence of FeOCl/H₂O₂ and BC/H₂O₂ systems were 70.58% and 0.76%, respectively, and the removal efficiencies of AO-II in the presence of the Fe₁C_X/H₂O₂ system ranged from 83.8% to 97.1%, which clearly shows that the introduction of BC to FeOCl significantly increased the removal rate of AO-II; this can be explained by the fact that BC provided numerous adsorption sites, which promoted the transfer of AO-II in solution to the catalyst surface, which could effectively decrease the migration distance of free radicals and promote the utilization efficiency of free radicals. Alternatively, BC has excellent electrical conductivity and plays the role of electron transport, which can reduce the recombination of e⁻-h⁺ pairs and enhance the carrier utilization efficiency, thus generating more reactive oxygen radicals. It was found that the removal efficiency of AO-II for Fe₁C_X composites first increases and then decreases with increasing BC content. When the ratio of FeOCl and BC was 1:0.2 (Fe₁C_0.2), the highest AO-II removal rate was achieved for Fe₁C_0.2 composite, and the removal efficiency of the AO-II reached 97.1% within 40 min. The reason may be that when the ratio of FeOCl to BC is from 1 to 0.2, the separation efficiency and photocatalytic activity of e⁻-h⁺ in FeOCl were improved.

Control experiments were performed under various conditions, including FeOCl + BC, no light, no catalyst, and no H₂O₂ (Figure S1). It can be seen from Figure S1 that the removal rate of AO-II for the Fe₁C_0.2 composite was higher than that of these control experiments. The degradation kinetics of AO-II followed pseudo-first-order behavior (Figure 1b), with rate constants (K) of 0.12623 min⁻¹ for Fe₁C_0.2, 0.04097 min⁻¹ for FeOCl, and 0.000203 min⁻¹ for BC. To further verify that the combination of FeOCl and BC has a synergistic effect in degrading AO-II, the synergistic factor of the final process following Equation (1) was used:

$$\mathrm{SF}={\displaystyle \frac{K_{FeOCl/BC}}{K_{FeOCl}+K_{BC}}}$$

(1)

where ${\mathrm{K}}_{FeOCl/BC}$, FeOCl, and ${\mathrm{K}}_{BC}$ are the kinetic constants of the Fe₁C_0.2 composite, FeOCl, and BC, respectively.

The SF value of the Fe₁C_0.2 composite was found to be 3.1 by calculation. This shows the synergy of FeOCl and BC in catalytic oxidation. The main reason may be that the BC substrate can boost the surface iron redox [26]. The photocatalytic activities of various Fenton-like catalysts studied in the literature are listed in Table S1 [23,32,33,34,35,36]. As a result, the photocatalytic activity of the as-obtained Fe₁C_0.2 composite is superior to that of most photocatalysts.

3.2. Sample Characterization

Figure 2a shows the XRD spectra of FeOCl, BC, and the Fe₁C_0.2 composite. The 2θ angles at 10.8°, 26°, 35.78°, and 38.23° are consistent with the crystal planes of FeOCl (JCPDS No: 74-1369) [22]. A wide peak was observed at 20–25°, showing the amorphous features of the original biochar structure [37]. In addition, the Fe₁C_0.2 composite has the characteristic peaks of FeOCl and BC, indicating that FeOCl is successfully loaded on BC.

The FTIR spectra of FeOCl, BC, and Fe₁C_0.2 composite are shown in Figure 2b. As shown in Figure 1, the peak at 677 cm⁻¹ corresponding to the Fe-O group indicates that the source of Fe-O is FeOCl [38], while the peak at 3400 cm⁻¹ corresponding to -OH groups is attributed to the adsorbed water molecules on the material surface. The peak at 1100 cm⁻¹ is associated with the CO- stretching vibration modes [39]. The C=C bonding of BC leads to a clear absorption band near 1587 cm⁻¹, indicating a high carbon content and strong aromatization in BC [30]. The stretching vibration at 2923 cm⁻¹ can be attributed to C-H bonding [6,40]. The FTIR spectrum of Fe₁C_0.2 composite displays the overlapping peaks of the two materials.

The surface characteristics of the prepared samples and their role in photocatalytic activity were examined through N₂ adsorption–desorption isotherms (Figure 2c). The presence of type V adsorption curves with H₃-type hysteresis loops in all materials confirms the existence of a hierarchical pore structure incorporating micropores, mesopores, and macropores [41]. The BET surface areas of FeOCl, BC, and the Fe₁C_0.2 composite were found to be 4.34, 1.73, and 17.3 m²·g⁻¹, respectively. The increase in the FeOCl/BC surface area indicated that the composite structure can provide more active sites compared to only FeOCl [42].

The SEM-EDX of FeOCl, BC, and the Fe₁C_0.2 composite are shown in Figure 3, and the microstructures of their samples are analyzed. FeOCl shows a more obvious lamellar structure. BC shows a porous structure, and the EDX image shows that BC contains a small amount of K and Si, probably due to its high content of alkali metals [43]. The SEM image of the Fe₁C_0.2 composite shows that FeOCl was grown on the surface of BC.

The UV-Vis DRS spectra of FeOCl, BC, and the FeOCl/BC composite are shown in Figure S2a,b. It can be seen from Figure S2a that FeOCl shows a clear absorption peak at 201 nm. However, the Fe₁C_0.2 composite exhibits a wide absorption edge at 300~400 nm, which can be attributed to the introduction of BC, showing its ability to extend the adsorption range. Concurrently, the optical band gap was calculated by analyzing Tauc plots following the relation $(\mathsf{\alpha }\mathrm{h}\mathsf{\nu }{)}^{{\displaystyle \frac{1}{\mathrm{n}}}}=\mathrm{A}(\mathrm{h}\mathsf{\nu }\text{-}{\mathrm{E}}_{\mathrm{g}})$ (Figure S2b). The forbidden band widths of FeOCl and the Fe₁C_0.2 composite are estimated to be 1.74 and 3.11 eV, respectively. Figure S2c shows the UV-Vis spectral changes of AO-II solution with reaction time in the FeOCl/BC composite/H₂O₂/Vis system. The maximum absorbance of AO-II quickly disappeared with reaction time in the range of visible wavelengths, indicating that the degradation process suffers a chemical reaction and that the structure of AO-II molecules has been destroyed rather than undergoing a simple physical adsorption process.

3.3. Effect of H₂O₂ Dosage and Fe₁C_0.2 Composite Concentrations

H₂O₂ reacts not only with Fe²⁺ to form •OH precursors but also with Fe³⁺ to regenerate Fe²⁺. To explore the effect of H₂O₂ consumption on the Fe₁C_0.2 composite Fenton system, four different H₂O₂ concentrations of 10, 30, 50, and 100 mmol were used. As shown in Figure 4a, as the concentration of H₂O₂ increased from 10 mmol to 100 mmol, the removal rate of AO-II decreased slightly from 98.2% to 95.1%, probably because the excess H₂O₂ would compete with the catalyst, the excess H₂O₂ would consume part of the generated •OH to form HOO•, which would also react with •OH to form water and oxygen [44,45,46], and the limited catalyst would not provide more active sites for H₂O₂ activation. Considering the balance between oxidative efficiency and reagent costs, 10 mmol H₂O₂ was selected as the standard condition for further experimentation.

To explore the effect of catalyst concentration on the Fe₁C_0.2/H₂O₂ Fenton system, four different catalyst concentrations of 60, 120, 240, and 360 mg·L⁻¹ were selected for investigation. As shown in Figure 4b, when the catalyst concentration was 60 mg·L⁻¹, the removal efficiency of AO-II reached 94.8%. As the catalyst concentration increased from 60 to 240 mg·L⁻¹, the removal efficiency of AO-II gradually increased first and then decreased; this is because as the catalyst concentration increases, it has more active sites for •OH generation. However, beyond a certain concentration threshold, catalyst particles tend to aggregate and precipitate from the solution.

3.4. Effect of the Initial pH Value

To evaluate how initial pH influences the Fenton reaction efficiency, the solution pH was precisely regulated via dropwise titration with 0.1 M H₂SO₄ or NaOH under continuous pH monitoring. Seven different initial pH values of 2.1, 3.1, 4.1, 6.1, 8.1, 9.1, and 11.1 were selected, and the degradation effect of AO-II was explored.

As shown in Figure S3, the catalytic activity was significantly low under extremely acidic and alkaline conditions, and the removal efficiency of AO-II was only 16.4% and 27.2% at pH = 2.1 and 11.1, respectively. Under acidic conditions (except pH = 2.1), the catalytic system had a very high removal rate. A 10% reduction in degradation efficiency was observed under basic conditions (pH < 11.1), presumably because leached Fe²⁺ ions precipitated as hydroxides, thereby becoming unavailable for H₂O₂ activation. The oxidative degradation of AO-II was considerable for a pH range from 3.1 to 9.1, where the removal rate of AO-II exceeded 90.8% in all cases. The results showed that the FeOCl/BC composite not only worked in a wide pH range but also maintained high catalytic activity across a wide pH range, suggesting good applicability.

3.5. Effect of Temperature

To investigate the effect of temperature on the Fe₁C_0.2/H₂O₂ Fenton-like system, three different temperatures of 15, 25, and 35 °C were selected. As shown in Figure S4a, the degradation rate of AO-II increased gradually with the increasing temperature. From Figure S4b, the primary kinetic rate constants of the photocatalytic removal of AO-II at 15, 25, and 35 °C are k₁ = 1.27615, k₂ = 0.97014, and k₃ = 0.84079, respectively. The 1/T~-lnk image can be plotted according to the Arrhenius equation. As shown in Figure S4c, the slope of the line is 1944.4, which is E_a/R. The activation energy of the reaction is calculated to be E_a = 16.2 kJ·mol⁻¹. The value of E_a in this study is slightly lower than that from the literature [46,47,48], showing that the energy needed for the degradation of AO-II by the Fe₁C_0.2/H₂O₂ Fenton-like system is relatively low and easy to attain.

3.6. Reusability of the Fe₁C_0.2 Composite and Versatility

The recyclability of catalysts represents a fundamental parameter for assessing their practical applicability. To evaluate the reusability of the Fe₁C_0.2 composite, the composite was recovered by centrifugation after each use and used in subsequent experiments. The removal rate of AO-II by the Fe₁C_0.2 composite after five reuses is displayed in Figure 5a. The removal rate of AO-II by the Fe₁C_0.2 composite decreased slightly with an increasing number of cycles. The removal efficiency of AO-II decreased only by 13.8% from 97.1% to 83.3% after five cycles; this indicated that the Fe₁C_0.2 composite was more reusable and had good prospects for practical applications. This decrease in the AO-II removal rate was mainly because of Fe leaching during the reaction. As shown in Figure 5b, the XRD pattern of the recycled Fe₁C_0.2 composite showed some change compared to those from the fresh Fe₁C_0.2 composite, which showed that the stability of Fe₁C_0.2 composite is average.

As shown in Figure 5c, when FeOCl and the Fe₁C_0.2 composite degraded AO-II in aqueous solution at pH = 3.1, the iron leaching concentrations were different, and the total and ferrous iron contents of the Fe₁C_0.2 composite were higher than those of FeOCl. As AO-II was degraded, the Fenton process continuously oxidized Fe²⁺ to Fe³⁺, resulting in progressively higher concentrations of aqueous iron species. When the degradation of AO-II was completed, the maximum leached Fe concentration of the Fe₁C_0.2 composite was 0.31 mg·L⁻¹; it accounted for only 0.49% of the total Fe in the catalyst, which was 0.12 mg·L⁻¹ higher than that of FeOCl. Therefore, FeOCl loading on biochar accelerated Fe³⁺ and Fe²⁺ conversion and enhanced the Fenton reaction.

Photo-Fenton degradation of other organic pollutants, including TCH, MB, and RhB (60 mg·L⁻¹), was used as simulated pollutants by the Fe₁C_0.2 composite (Figure 5d). The degradation efficiency of TCH, MB, and RhB reached 91.6%, 95.0%, and 99.3% at 40 min, respectively. The photodegradation experiments suggested that the Fe₁C_0.2 composite could find practical application for photo-Fenton degradation of various organic pollutants.

3.7. Effect of Coexisting Anions

Various inorganic anions (Cl⁻, NO₃⁻, and SO₄²⁻) and NOM (humid acid referred to as NOM in this study) existing in wastewater could affect the photo-Fenton degradation of organic pollutants. It can be seen from Figure S5a that the inhibitory effects of different anions (10 mM) on the oxidative ability were in the order of Cl⁻ < SO₄²⁻ < NO₃⁻. The oxidative ability of chlorine radicals (Cl) generated by the reaction of ·OH with Cl⁻ is weak [49,50]. The results showed that Cl⁻ had a slight inhibitory effect on the catalytic ability. The introduction of NO₃⁻ reduced the removal efficiency slightly due to its tendency to react with h⁺ and OH [51]. Moreover, NO₃⁻ can react with e⁻, decreasing the combination of h⁺ and e⁻ [52]. SO₄²⁻ showed a suppressive effect because it captured OH and h⁺ [48], contributing to inhibition of AO-II degradation. NOM had an inhibitory influence on the degradation of AO-II. The reason may be that NOM competes with the photocatalyst and AO-II for photons and reactive oxygen species [53]. It is generally believed that cations should not affect the degradation of AO-II because their state is stable. Here, the removal of AO-II was explored under the conditions of NaNO₃, Mg(NO₃)₂, and Ca(NO₃)₂. It can be seen from Figure S5b that the three salts exhibit inhibitory effects on AO-II removal. The reason may be ascribed to nitrate ions present in solution.

3.8. Active Species Involved in the Fe₁C_0.2/H₂O₂ Fenton System

To elucidate the effect of reactive additions on the Fe₁C_0.2/H₂O₂ Fenton system, a sequence of reactive species capture experiments was conducted to discern the reactive radical species involved in the reactions. IPA (•OH scavenger), BQ (•O₂⁻ scavenger), methanol (h⁺ scavenger), and AgNO₃ (e⁻ scavenger) were added to the AO-II to be tested [54]. KI was used as a scavenger of •OH generated on the catalyst surface. As seen in Figure 6a, the effect of BQ was less, and the degradation efficiency of AO-II was still 94.6%, which showed that the removal of AO-II from aqueous solution by •O₂⁻ did not play a major role. In contrast, IPA, methanol, and AgNO₃ had greater effects, and the degradation efficiencies of AO-II were 27.3%, 57.4%, and 51.1%, respectively. Selective scavenging with 10 mM KI (specific for surface •OH) was employed to discriminate between interfacial and bulk-phase •OH contributions. These results indicated that •OH was produced both on the catalyst surface and in solution. AgNO₃ showed high inhibition as an electron scavenger. Therefore, we speculated that the degradation of contaminants by the Fe₁C_0.2 composite was mainly due to OH, while •O₂⁻ and h⁺ played minor roles.

EPR spectroscopy was employed to identify the active species in the system. No signals for DMPO-•O₂⁻ and DMPO-•OH adducts were found under dark conditions (Figure 6b,c). The typical 1:2:2:1 quartet peak of the DMPO-•OH signal can be observed under the irradiation of simulated sunlight. Characteristic peaks with a relative intensity of approximately 1:1:1 were detected, showing the presence of •O₂⁻. The signal intensity of DMPO-•OH is slightly higher than that of DMPO-•O₂⁻. This fact showed the OH is more important for AO-II removal than the •O₂⁻ free radicals. Under visible light radiation, photoinduced h⁺/e⁻ pairs were produced by the Fe₁C_0.2 composite, and the photoinduced e⁻ reacted with O₂ to form •O₂⁻ radicals. The photoinduced holes (h⁺) attenuated the TEMPO signal via formation of a diamagnetic TEMPO-h⁺ spin adduct, which lacks the unpaired electrons required for EPR detection. Therefore, the signal weakened under the irradiation of simulated sunlight, showing the production of h⁺ (Figure 6d).

To obtain further insight into the TCH degradation mechanisms by the Fe₁C_0.2 composite, XPS investigation of the Fe₁C_0.2 composite before and after the reaction was performed. As expected in Figure 7a, the full-scan spectrum illustrates the coexistence of Fe, O, Cl, and C within the Fe₁C_0.2 composite. The chemical composition analysis based on the full XPS survey revealed a lower C content (44.99 wt%) in the Fe₁C_0.2 composite after the reaction. As seen in Figure 7b, the peaks at 711.26 and 724.32 eV are attributed to Fe 2p3/2 and Fe 2p1/2, respectively, showing that the Fe of FeOCl is present in the Fe (III) state [55]. The Fe 2p3/2 and Fe 2p1/2 peaks in the Fe₁C_0.2 composite exhibit a negative shift to 710.59 eV and 723.71 eV, respectively, compared to pristine FeOCl. This 0.67 eV reduction in binding energy indicates electron density redistribution at the FeOCl-BC interface, likely through π-d orbital hybridization between BC’s graphitic carbon and Fe³⁺ centers. Compared with the fresh Fe₁C_0.2 composite, the peak ratio of Fe (III) increased from 39.92% to 42.86%, suggesting that some of the Fe (II) was oxidized to Fe (III), which could be attributed to the interaction with H₂O₂ producing •OH. It can be seen from Figure 7c that the C1s spectrum of BC is fitted to three peaks at 283.98, 284.56, and 285.25 eV. The C 1s spectrum of the Fe₁C_0.2 composite can be fitted to three peaks at 284.38, 285.28, and 288.22 eV, which correspond to C-C, CO, and CO-= bonds [28], respectively, and the peak phase is shifted in the direction of high binding energy, showing that there is a strong interaction between FeOCl and BC. Compared with the fresh Fe₁C_0.2 composite, the peak ratio of CO-= increased from 36.88% to 56.61%, while the proportion of C-O groups decreased from 52.88% to 13.47%. All these results suggested that the C-O group was oxidized to CO-= when H₂O₂ was reduced to H₂O since C-O groups could be electron-donating moieties in biochar [55]. The Cl 2p peak at 199.21 is attributed to Fe-Cl in FeOCl in Figure 7d.

3.9. Possible Photocatalytic Principles of Catalysts

The mineralization efficiency of the Fe₁C_0.2/H₂O₂/vis system was determined. The total organic carbon (TOC) removal efficiency for TCH by the Fe₁C_0.2/H₂O₂/vis system reached a value of 43% within 30.0 min, slightly higher than that of the FeOCl/H₂O₂/vis system (Figure S6).

Based on the experimental findings outlined in previous sections, Figure 8 illustrates the proposed reaction mechanism through which the Fe₁C_0.2 composite facilitates the degradation of AO-II molecules. Under visible light excitation, the valence band electrons leap to the conduction band, forming e⁻ and h⁺ (Equation (2)), and the photogenerated e⁻ in the conduction band reduces O₂ to produce •O₂⁻ (Equation (3)). The Fe³⁺/Fe²⁺ redox cycle is extended because the active functional groups in the biochar matrix, such as sp² C = C groups, allow Fe³⁺ to be reduced by transferring e⁻ to regenerate the Fe²⁺ species (Equations (4) and (5)) to ensure effective •OH production and subsequent AO-II degradation. The active •OH produced from ≡Fe²⁺ and H₂O₂ was easily facilitated by the Fe₁C_0.2 composite (Equation (6)). ${\mathrm{Fe}}^{3+}$ can in turn produce •O₂⁻ by a Fenton-like reaction mechanism to form a cyclic mechanism with ${\mathrm{Fe}}^{2+}$ (Equation (7)). Highly active OH may initially attack the structural center of AO-II, which would cause additional ring opening and mineralization. •OH and other reactive species, such as h+ and •O₂⁻, are all involved in AO-II degradation (Equation (8)).

$${\mathrm{Fe}}_1{\mathrm{C}}_{0.2}\stackrel{\mathrm{h}\mathsf{\nu }}{\to }{\mathrm{F}}_1{\mathrm{C}}_{0.2}(h^{+}+e^{-})$$

(2)

O₂ +e⁻_CB →•O₂⁻

(3)

BC → e⁻

(4)

e⁻ + Fe³⁺→Fe²⁺

(5)

$${\mathrm{Fe}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{3+}+{\mathrm{OH}}^{-}+•\mathrm{OH}$$

(6)

$${\mathrm{Fe}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{2+}+2{\mathrm{H}}^{+}+•{\mathrm{O}}_2^{-}$$

(7)

•OH/h⁺/•O₂⁻ + AO-II →CO₂ + H₂O + intermediates

(8)

4. Conclusions

In this study, four Fe₁C_X (X = 0.1, 0.2, 0.5, and 1.0) composites were prepared by simple calcination using bamboo powder biochar as a carrier. Fe₁C_X composites were characterized and used as catalysts for the photocatalytic degradation of AO-II. These results displayed that among the Fe₁C_X composites, the Fe₁C_0.2 composite had the highest ability to degrade AO-II and showed a good ability to degrade AO-II under acidic and weakly basic conditions. These results show that there are synergistic interactions between FeOCl and biochar. The capture experiment showed that the removal of AO-II by the Fe₁C_0.2 composite was mainly dominated by •OH; however, •O₂⁻ and h⁺ played a minor role. The removal efficiency of AO-II by the Fe₁C_0.2 composite was only reduced by 13.8% after five repeated uses. In addition, the Fe₁C_0.2 composite showed a good removal ability for different organic pollutants. The results obtained in this study provide a new idea for the development of FeOCl-based photocatalysts in photo-Fenton wastewater treatment.