Fe@Fe₂O₃-loaded biochar as an efficient heterogeneous Fenton catalyst for organic pollutants removal

Various organic pollutants, such as aromatic pollutants, are often detected in wastewater effluents of pharmaceutical manufacturers and surface water. Aromatic pollutants are resistant to biodegradation or other methods, because of the stable benzene ring structure. In addition, the toxicity of aromatic pollutants can cause serious harm to the environment and human health (Ye et al. 2020). In recent years, demand for all forms of chemical raw materials has increased with economic development. As a result, sudden water pollution incidents are likely to occur more frequently. For example, phenol leakage occurred in 2020 in Guangdong Province, China and the leaked phenol flowed into the Rong River, resulting in extremely serious environmental pollution. Incidents like this are often characterized by high concentration and large scale, and reducing their harmfulness to the minimum range efficiently without new environmental pollutants is challenging.

Advanced oxidation processes (AOP_S) have been regarded as an efficient method to remove organic pollutants from wastewater (Bilińska & Gmurek 2021; Rayaroth et al. 2021; Lin et al. 2022; Yuan et al. 2022), including traditional Fenton oxidation, electrochemical oxidation (Yang et al. 2021), ozonation (Xu et al. 2022), photochemical oxidation (Kilic et al. 2019) and microwave oxidation (Adekunle et al. 2021), etc. Fenton oxidation is better-suited to deal with sudden environmental water pollution incidents compared with other AOPs, because other AOPs involve complex operation and time-consuming, special devices. Therein, persulfate oxidation has received more attention because of its excellent catalytic performance (Su et al. 2020; Zhang et al. 2022). However, the Fenton reaction is an efficient method to deal with sudden water pollution events as there is no secondary pollution (Xiaoliang et al. 2021).

At present, the most commonly used homogeneous Fenton method has the disadvantage of producing a large amount of iron sludge and narrow pH ranges (Wang et al. 2016). Therefore, many studies have focused on the heterogeneous catalyst to alleviate the formation of iron sludge (Xu & Wang 2011; Zha et al. 2014). Among numerous heterogeneous catalysts, Fe and its oxides nanoparticles have received more attention due to its abundant reserves and excellent catalytic performance. However, nanoparticles can gradually agglomerate during the application process, resulting in a sharp reduction in the ability to activate H₂O₂. Hence, a suitable matrix is often adopted to effectively load the metal or its oxide, so as to improve the ability to activate H₂O₂ (Pan et al. 2021). However, the reaction time for heterogeneous Fenton reaction based on Fe or its oxide is generally about 20–60 min. Sudden water pollution events require rapid and effective treatment, so finding a cost-effective and efficient heterogeneous Fenton catalyst to treat these events is of great importance.

Biochar is often used as a matrix of various composite materials because of its abundant pores and large specific surface area (Liang et al. 2021; Qi et al. 2022). In this study, the discarded tea stalks were recycled into composites materials by iron divalent solution dipping and subsequent calcination, using a heterogeneous Fenton catalyst to quickly treat phenol pollution. This study will explore new possibilities of using organic solid waste to synthesise composite materials to treat wastewater.

Ferrous sulfate (FeSO₄·7H₂O, 99%), sodium borohydride (NaBH₄, 98%), sulfuric acid (H₂SO₄, 98%), sodium hydroxide (NaOH, 96%), anhydrous ethanol (C₂H₆O, 99%), phenol (C₆H₅OH, 99%), ciprofloxacin (C₁₇H₁₈FN₃O₃, 98%), enrofloxacin (C₁₉H₂₂FN₃-O, 98%), norfloxacin (C₁₆H₁₈FN₃O₃, 98%), bisphenol A (C₁₅H₁₆O₂, 99.99%) and methanol (CH₄O, 99.99%) were purchased from Sinopharmaceutical Chemical Reagent (China) Co., Ltd. Hydrogen peroxide (H₂O₂, AR, 30%), benzoic acid (C₇H₆O₂, 99%), ammonium ferrous sulfate (Fe(NH₄)₂(SO₄)₂6H₂O, 99%), hydroxylamine hydrochloride (NH₃OHCl, 99%), 1meme 10-phenanthroline (anhydrous) (C₁₂H₈N₂, 99%), and tert-butanol (C₄H₁₀O, 99%) were purchased from China Aladdin Chemical Co., Ltd. All reagents were used directly without further purification, and the aqueous solutions involved in the experiment were all prepared with ultra-pure water.

Biochar was produced by the pyrolysis of tea stalk waste (sourced from FuJian Province, China). First, the collected solid was washed with ultra-pure water three times to remove surface dirt, and then dried in an oven at 60 °C. Then, the dried tea stalk waste was crushed and sieved (60-mesh sieve) into powder, which was pyrolyzed for 4 h at 600 °C (heating rate 5 °C/min) under Ar (purity 99%) atmosphere. The biochar collected was named BC.

1 g of biochar (BC) was dispersed in 50 mL of Fe²⁺ solution (0.28 mol/L), and the suspension was placed on a shaker operated at 200 rpm for 8 h. The precipitate was collected by centrifugal separation and washed with ultra-pure water and ethanol three times. The solid obtained was dispersed in 100 mL of ultra-pure water, with 40 mL of sodium borohydride (0.08 mol/L) slowly dripped in, then followed by reaction for 1 h. Finally, the black precipitate was washed with ultra-pure water and ethanol three times and dried in the oven at 60 °C (named precursor). The powders were calcined under Ar (purity 99%) atmosphere at various temperatures (600, 700, 800 and 900 °C) at a heating rate of 5 °C/min, and kept for 1 h. The sample obtained was recorded as FSC-X00 (X is the calcination temperature), and was stored in a vacuum dryer for later experiments.

The crystal structure information of FSC-X00 was obtained by X-ray diffractometer (XRD, Rigaku MiniFlex 600, Japan). The morphology and structure of FSC-600 were observed by scanning electron microscope (SEM, FEI QUANTA 250, USA). The functional group information of FSC-X00 was collected by Fourier transform infrared spectroscopy (FT-IR, Thermo Scientific Nicolet iS10, USA). The composition and chemical structure of FSC-600 were analyzed by Thermo Fisher Scientific K-Alpha X-ray spectroscopy (XPS, Thermo Scientific ESCALAB 250Xi, USA). The binding energies of all elements were calibrated with reference to C1 s (284.8 eV).

All experiments, unless otherwise specified, were conducted in a 100 mL glass flask in a constant temperature shaker (25 °C). Typically, a 20 mg catalyst was first dispersed into 50 mL of phenol aqueous solution (20 mg/L). Solution pH was adjusted to 3, 4, 6, 8 with NaOH (1 mol/L) or H₂SO₄ (1 mol/L) solution.

The concentration of phenol was tested by UV-vis spectrophotometry. The concentrations of CIP, ENR, NOR and BPA were analyzed by High Performance Liquid Chromatography (HPLC, Agilent 1200, USA). The separation of degradation products was conducted by Agilent Eclipse XDB-C18 column (5 μm × 4.6 mm × 250 mm). Mobile phase A was methanol, mobile phase B was 0.1% formic acid and water mixed solution (70 : 30 v/v), the flow rate was 0.5 mL/min, the injection volume was 20 μL, and the detection wavelengths were 279 nm (CIP (Normile et al. 2017)), 290 nm (BPA (Yang et al. 2018; Zhang et al. 2021)), 278 nm (ENR (Dror et al. 2020; Xiao et al. 2020)) and 276 nm (NOR (Chen et al. 2021; Mohan et al. 2021)). Samples were collected at 30 min, 60 min, 120 min, 240 min, 480 min and a total organic carbon analyzer (TOC, Analytik Jena AG, Germany) was used to assess the changes in organic carbon.

The concentrations of Fe²⁺ and total irons were measured by a 1,10-phenanthroline method. Samples were analyzed by measuring the characteristic absorption peak of Fe²⁺-1,10-phenanthroline complex at 510 nm with a UV–vis spectrophotometer. Hydroxylamine hydrochloride was used as the reducing agent for the total iron concentration measurement.

Different dosages of H₂O₂ (2 mmol/L, 4 mmol/L, 8 mmol/L, 20 mmol/L) and different dosages of FSC-X00 (0.1 g/L, 0.2 g/L, 0.3 g/L, 0.4 g/L) were added into the phenol solution and stirred. To compare the role of BC, FSC-X00, H₂O₂ during the degradation process, five groups of experiments (BC, BC/H₂O₂, H₂O₂, FSC-X00, FSC-X00-H₂O₂) were conducted. After washing and drying, the used FSC-600 was tested for cycle stability and reactivated by high-temperature calcination named FSC-600H.

At each interval (0.5, 1, 2, 5, 10, 15, 30, 40 min), 2 mL of solution was removed from the reactor, and filtered by a 0.22 μm pore film to measure the concentration of phenol. All the above experiments were repeated three times.

Figure 1(a) and 1(b) show the XRD patterns and FTIR spectra of FSC-X00 materials. As seen in Figure 1(a), the diffraction peak at 35.4 °, 45.1 ° and 65.02 ° can be attributed to the (311), (110) and (200) crystal plane of Fe₂O₃ (PDF#73-0603). The peaks locating at 45.1 ° and 65.02 ° are attributed to Fe⁰ (PDF#87-0721). But the peak at 26.2 ° attributed to biochar (PDF#75-0444, Figure S1a) is invisible because the peak intensity of Fe₂O₃ and Fe⁰ outperform that of biochar. In addition, the intensity of the diffraction peak of Fe⁰ increases with the increase of calcination temperature. However, the peak intensity of Fe₂O₃ increases first and then decreases with the increase of calcination temperature. When the calcination temperature is 900 °C, the diffraction peak of Fe₂O₃ disappears. We think that Fe₂O₃ particles underwent grain growth before 700 °C and Fe₂O₃ was reduced to Fe⁰ by biochar after 700 °C, which is consistent with previous research (Dong et al. 2016). From Figure 2(b), we can see that the Fe-O stretching vibration band belonging to Fe₂O₃ at 681 cm⁻¹ gradually disappears when the temperature increases from 600 °C to 900 °C, which is consistent with the XRD analysis. Li et al. (Li et al. 2020) found a similar experimental phenomenon. Figure S1b shows that the peak of precursor at 1,105 cm⁻¹ can be attributed to the stretching vibration band in C-OH, and its band shifts to 1,232 cm⁻¹ after calcination, which may be due to the transformation of C-OH to C-O-C during calcination. The peak gradually disappears when the calcination temperature continues to rise at 800 °C, which may be due to bond-breaking as temperature rises (Liu et al. 2010).

Figure 2 displays the SEM images of biochar BC-600 and FSC-600. As shown in Figure 2(a) and Figure S2a, the biochar exhibits abundant honeycomb-like macropores and the BET specific surface areas (SSA) and average pore size (APS) of BC were 149 m²/g and 11.69 nm (Table S1), respectively, which are beneficial for the flow of solution, enabling the loading of Fe²⁺. In addition, the surface negative charge promoted the adsorption of Fe²⁺ (Figure S2b), ensuring the formation of Fe or Fe₂O₃. As shown in Figure 2(b), the tubular pores were filled with a large number of particles after calcination, which can be attributed to Fe and Fe₂O₃ according to XRD data.

Figure 3 shows XPS spectra of FSC-600. Figure 3(a) shows that the peak at 710, 532, and 284 eV correspond to Fe 2p, O 1 s, and C 1 s, respectively. This proves that the catalyst was composed of Fe, O and C element. Figure 3(b) shows the Fe 2p high-resolution spectra of FSC-600 before and after the reaction. The peaks located at 710.8 eV and 724.6 eV correspond to 2p_3/2 and 2p_1/2 of Fe²⁺, and the peaks at 714.4 eV and 727.4 eV are assigned to the 2p_3/2 and 2p_1/2 of Fe³⁺, indicating the existence of Fe²⁺ and Fe³⁺ in FSC-600 (Wu et al. 2015). In addition, the satellite peak of Fe(III) was observed at 719.62 eV in FSC-600 (Grosvenor et al. 2004), further proving the existence of Fe₂O₃ in FSC-600. According to the XRD characterization, Fe₂O₃ and Fe⁰ coexisted in FSC-600. However, because of the limited penetration depth of XPS (Lin & Chen 2017), no Fe⁰ signal appeared in FSC-600.

The pH value plays an important role in Fenton reaction and degradation of organic contaminants, so the degradation performance of materials obtained at different calcination temperatures were evaluated under different pH values. As shown in Figure 4, FSC-X00 showed good degradation performance when the pH value was 3, the degradation efficiency of phenol was close to 100% after 5 min except for FSC-900. However, phenol can still be mostly degraded after 15 min by FSC-900. When the pH value was 4, FSC-X00 material also showed relatively good degradation efficiency and reached 100% after 15 min. However, as the pH value increased, the phenol degradation efficiency of FSC-X00 decreased. The degradation efficiency of FSC-X00 in 40 min decreased from 100% to 40% when the pH value was 6. When the pH value rose to 8, the degradation efficiency decreased more significantly and the degradation efficiency of FSC-X00 decreased to about 15%. This suggests that acidic solutions are more favorable for the degradation of phenol by FSC-X00-H₂O₂ system. Possible reasons include: (1) When the solution is acidic, Fe⁰ can easily enter the solution to form Fe²⁺, thus promoting H₂O₂ to produce •OH; (2) •OH has a higher reduction potential than other free radicals under acidic conditions (Mao et al. 2019; Fang et al. 2021); (3) Under acidic conditions, it is easier for Fe²⁺ to form Fe(OH)₃; (4) Excess OH⁻ will react with H₂O₂ to generate O₂ under alkaline conditions, reducing the degradation efficiency (Dong et al. 2016). Figure 4 shows that the degradation efficiency of the samples obtained at different calcination temperatures did not differ significantly when the pH value was 3 or 4, so a pH value of 3 was selected for the following experiments. In addition, from the perspective of energy consumption, FSC-600 was selected as the subsequent experimental material.

To obtain the best experimental parameters, the dosage of H₂O₂ and catalyst were screened respectively. As shown in Figure 5(a), the degradation efficiency in 40 min was about 90% when the H₂O₂ concentration was 2 mmol/L. It may be that the H₂O₂ concentration was insufficient and the •OH generated failed to complete the degradation of all phenol. The degradation efficiency was significantly enhanced when the H₂O₂ concentration was increased from 2 mmol/L (70%) to 4 mmol/L and 8 mmol/L (nearly 100%) in 5 min. However, the degradation efficiency had decreased by 10% in first 5 min when the H₂O₂ concentration reached 20 mmol/L. This may be due to the production of excessive •OH reacting with H₂O₂ to produce weak oxidizing peroxy radicals (Deng et al. 2018), causing degradation efficiency to decrease at the beginning (•OH+ H₂O₂→ HO₂• + H₂O). Therefore, excessive H₂O₂ may reduce the utilization rate of •OH, thereby impacting the degradation efficiency of phenol. Considering that the effect of H₂O₂ concentration is similar between 4 mmol/L and 8 mmol/L, 4 mmol/L was chosen to conduct the subsequent experiments.

From Figure 5(b), it is not difficult to find that only 30% of phenol was degraded within 40 min when the dosage was 0.1 g/L. The degradation efficiency of phenol increased from 30% to 80% (40 min) and 100% (15 min) when FSC-600 were 0.2 g/L and 0.3 g/L, respectively. The degradation efficiency could reach nearly 100% in 5 min with the dosage of FSC-600 raised up to 0.4 g/L. Therefore, 0.4 g/L was selected as the most suitable dosage.

The degradation efficiency of FSC-600-H₂O₂ systems is shown in Figure 6(a) with the BC, BC-H₂O₂, H₂O₂, and FSC-600 as control. When H₂O₂ was used alone, less than 1% of phenol was degraded within 40 min, which indicated it had almost no effect on phenol degradation. 10% of phenol was degraded by BC, BC-H₂O₂ within 40 min, which suggested that BC had no obvious activation effect on H₂O₂. When FSC-600 was used without H₂O₂, the removal efficiency of phenol was only 20%, which may come from the adsorption. However, phenol was almost completely degraded in FSC-600-H₂O₂ system within 5 min, which indicated that FSC-600 had an obvious activation effect on H₂O₂. The pseudo-first-order kinetics model was used to fit the kinetic process (Bhattacharya & Mazumder 2020; Duan et al. 2021). As shown in Figure S3, with the increase of phenol concentration, the degradation efficiency gradually decreased and the data fit well with pseudo-first-order kinetics model. Therefore, we believe that the degradation reaction is still dominated by kinetic control. Obviously, the k value for 50 mg/L phenol was the highest, showing that the FSC-600-H₂O₂ system has an quick kinetic process for the phenol with low concentrations. In addition, Figure 6(b) shows the degradation efficiency of FSC-600-H₂O₂ system for different kinds of organic compounds. FSC-600-H₂O₂ system also shows a good degradation efficiency on CIP, BPA and NOR and the degradation efficiency can reach 100% within 5 min. The degradation efficiency of ENR was slightly worse than that of the BPA, CIP and NOR, but can still reach 95% within 5 min. When the time was extended to 40 min, the degradation efficiency was close to 97%. This shows that FSC-600 is an excellent H₂O₂ activator, and has a strong degradation effect on a variety of new pollutants within a short time period.

For advanced oxidation technology, whether organic pollutants are completely degraded into inorganic carbon is also an important indicator for evaluating performance. Hence, the total organic carbon analyzer (TOC) and three-dimensional fluorescence analyzer were used to evaluate the degree of degradation by the FSC-600-H₂O₂ system. As shown in Figure S4, about 80% of the organic carbon was converted into inorganic carbon within 30 min, and the equilibrium was reached in 60 min (about 85%). Figure 7 shows the EEM diagrams of the reaction solution at 0 and 30 min. The fluorescence center of the initial solution is in area I (Ex = 270–350 nm/Em = 240–280 nm), which corresponds to the aromatic protein in phenol. After reacting for 30 min, the fluorescence intensity weakened and the fluorescence center shifted to area II, indicating that most of the phenol was gradually decomposed into small molecular substances within 30 min. As shown in Figure S5, phenol, succinic acid, and oxalic acid were found during phenol degradation from the initial 5 to 30 min, but phenol was completely degraded after 30 min. It is well-known that •OH can attack the benzene ring by electrophilic addition reaction. Therefore, the degradation pathways of phenol were speculated. Firstly, the benzene ring was attacked to form succinic acid and then the structure was further converted into chain-like molecules of intermediates, such as oxalic acid, by oxidizing species. Finally, chain-like molecules of intermediates were mineralized into CO₂ and H₂O•. At present, the mineralization effect of iron-based materials on organic pollutants is about 70–80%, and the mineralization time is longer than that of our work (Table 1), which shows that the FSC-600 is a promising heterogeneous Fenton catalyst for degrading organic pollution within a short timeframe.

To study the availability of the FSC-600-H₂O₂ system in real wastewaters, we simulated real waters by introducing natural organic matter (fulvic acid, FA) or inorganic ions (Na₂SO₄) into the phenol solution. Figure 8(a) shows the degradation efficiency of phenol in the FSC-600-H₂O₂ system under different concentrations of FA. First, when the concentration of FA was 10 mg/L, the degradation efficiency was close to 100% within 5 min. In other words, the degradation efficiency was almost unaffected. However, when the concentration of FA was 20 mg/L, the degradation efficiency reduced by about 10% in the first 15 min, and it could still be basically degraded after 40 min. When the concentration of FA rose to 40 mg/L, the degradation efficiency decreased by about 20% within 40 min. The decrease may be due to the following reasons: (1) The active site was occupied by FA molecules (Mady et al. 2019), which lowered the adsorption of phenol; (2) Part of •OH degraded FA instead of phenol. Although the degradation efficiency decreases with the increase of FA concentration, FSC-600 can still degrade most of phenol in a short time. In addition, we investigated the degradation efficiency in the presence of inorganic salt. Figure 8(b) shows that when the Na⁺ concentration was 1–2%, the FSC-600-H₂O₂ system can degrade nearly 100% of phenol within 5 min. However, the degradation efficiency decreased by about 10% within 5 min and it can still attain nearly 100% within 15 min when Na⁺ concentration was 3%. Interestingly, the degradation efficiency appeared to vary mildly in Min River and Seawater (Figure S6). Therefore, the competition experiments indicate that FSC-600-H₂O₂-system may be promising not only in real water bodies but also in the domain of biphasic media (Madriz et al. 2011).

The cycle stability of the catalyst is an important indicator for evaluating whether it can be applied in practice. Therefore, cycle stability experiments were carried out using FSC-600. As indicated in Figure 9(a), when the first recycled FSC-600 was used to conduct degradation experiments, the degradation efficiency decreased from 100% to about 90% within 40 min. However, the degradation efficiency was only about 60% in the second run. To understand why degradation efficiency decreased, a phase analysis of the recycled FSC-600 was conducted. As shown in Figure 9(b), the diffraction peaks of Fe⁰ for the recycled FSC-600 greatly decreased after the experiment, indicating that Fe⁰ was gradually consumed during the experiments. However, the diffraction peak of Fe₂O₃ still exists after two rounds of experiments. Therefore, the FSC-600 was reactivated by high-temperature calcination after two rounds of experiments (named as FSC-600H). As shown in Figure 9, the FSC-600H has a strong Fe⁰ diffraction peak and its degraded efficiency towards phenol reached nearly 100% within 15 min. This means that FSC-600 can be activated for secondary use by simple calcination.

In this study, FSC-X00 was prepared from tea stalks waste and used as a heterogeneous Fenton catalyst for organic pollution degradation within a short time period. The optimum conditions were FSC-600 = 0.4 g/L, pH = 3, H₂O₂ = 4 mmol/L. The degradation experiments showed that FSC-600 could degrade phenol, CIP, BPA and NOR with nearly 100% within 5 min. The competition experiments demonstrated that the FSC-600-H₂O₂-system presented superior anti-interference performance. EPR spectra analysis showed that •OH deriving from the activation of H₂O₂ by Fe²⁺ played a dominant role in the oxidation process. In addition, a part of ¹O₂ resulting from O₂ was also involved during the degradation of phenol. Finally, FSC-600 can be recycled by annealing. This work indicates that the FSC-600-H₂O₂-system may be promising not only in real water bodies but also in the domain of biphasic media, and provides a reference scheme for using organic solid waste to prepare heterogeneous Fenton catalyst.

Diwei Chen: Conceptualization, Methodology, Data Curation, Investigation, Formal analysis, Writing – Original Draft, Visualization; Yonghao Wang: Conceptualization, Methodology, Validation, Writing – Review & Editing, Supervision; Zhiyan Zheng: Resources, Investigation; Feiji Zhang & Rufu Ke & Nan Sun: Software; Yongjing Wang: Funding acquisition

The authors would like to acknowledge the National Natural Science Foundation of China (No. 52000035) and the National Science Foundation of Fujian Province (No. 2020Y0016).

All relevant data are included in the paper or its Supplementary Information.