S-doping Fe–Ce composites derived from PBA to accelerate Fe(III)/Fe(II) cycle in the Fenton-like process

Abstract

The design of metal–organic framework-derived metal oxides/sulfides is an effective approach for water remediation. A series of Fe–Ce composites (Fe–Ce–S–N) with cluster-like morphology were prepared with prussian blue analogue (PBA) as the precursor in a hydrothermal process. S doping promoted the conversion of Fe(III) to Fe(II), which greatly improved the performance of the Fenton-like reaction. The optimized Fe–Ce–S–N catalyst exhibited superior catalytic activity towards sulfamethoxazole (SMZ) degradation (18.63 mg_SMZ g_cat⁻¹ min⁻¹), showing the largest k value (0.1862 min⁻¹), which was 2.1 and 8.1 times higher than that of Fe–Ce–S–H and Fe–Ce–N, respectively. Moreover, the Fe–Ce–S–N catalyst displayed a remarkable reusability, stability and adaptability in Fenton-like systems. Overall, this work provides new insights into extending novel PBA-derived catalysts to significantly improve heterogeneous Fenton-like performances.

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Environmental significance

Facilitating the reductive regeneration of catalyst metal centers to maintain a stable high catalytic activity is quite a challenge for most Fenton-like processes. The synthesized Fe–Ce–S–N had a three-dimensional cluster-like morphology that favors the exposure of more active sites, while the S doping facilitated the acceleration of the electron transfer in the Fenton process and promoted the conversion of Fe(III) to Fe(II), which greatly improved the performance of the Fenton-like reaction. In addition, the Fe–Ce–S–N/H₂O₂ system showed excellent catalytic activity and stability over a wide pH range and in the presence of coexisting substances. In conclusion, our work provided a new strategy for designing PBA-derived catalysts for wastewater treatment.

1. Introduction

Sulfamethoxazole (SMZ) is a sulfonamide antibiotic widely used in human and animal medicine.¹ However, due to its limited bioavailability, over 70% of SMZ is excreted into the natural environment through urine and feces. The discharge of SMZ into aquatic environments poses a significant threat to both human health and environmental safety, given its non-volatility and poor biodegradability.^2,3 Therefore, there is a high demand for an efficient method to remove SMZ from water.

Notably, since Fenton found that iron could activate H₂O₂ to degrade tartaric acid (TA), Fenton and Fenton-like reactions have drawn remarkable attention because of their facile operation, no requirement for extra energy and low cost of H₂O₂.^4–6 However, the disadvantages of homogeneous Fenton reaction including acidic reaction conditions, large consumption of ferrous salts and oxidants, unsatisfied kinetics of iron cycle and iron sludge generation have limited its application. In comparison, the heterogeneous Fenton process has emerged as a prominent research direction in recent years. Nevertheless, the heterogeneous Fenton efficiencies were not desirable because of several inherent shortcomings, such as the insufficient active sites and low H₂O₂ activation efficiency.^7–9

In this case, the primary concern for heterogeneous Fenton technology lies in the development of a catalyst that boasts abundant active sites and high efficiency in generating active species. To this end, researchers have successfully developed an array of novel catalysts by regulating their morphology, composition, facets and surface sites to activate H₂O₂ into free radicals for organic pollutant removal.^10–12 Among these catalysts, Fe-based oxide/sulfide was commonly employed for activating H₂O₂ due to the intricate structure and variable valence. However, their catalytic abilities may be limited by the inferior stability and low efficiency of the iron cycle.^13–15 Therefore, it is crucial to expedite the Fe(II)/Fe(III) cycle and enhance catalyst stability. Constructing bimetallic oxide/sulfide not only possesses a more diverse composition and valence states, leading to a variety of redox reactions, but also can significantly improve the catalytic activity and material stability. Prussian blue analogue (PBA) has the coordination framework structure with cyano-linked bimetallic centers, which can be employed as the desirable template for preparing bimetallic oxide/sulfide due to the ordered crystal structure, large surface area and tunability.^16–18

Based on of previous work,¹⁹ we found that the partial replacement of Fe with Ce could greatly improve the adsorption affinity of the catalyst for H₂O₂. However, the obtained Fe–Ce-PBA could not maintain the high catalytic efficiency due to leakage. Inspired by the aforementioned analysis, a novel Fe–Ce composite (composed of FeS, CeS and FeCe₂O₄) was synthesized through facile sulfurization of PBA using the hydrothermal process. Then, the effects of different reaction conditions in Fenton-like system were thorough investigated. Furthermore, the stability of the catalyst was verified and the degradation pathways and mechanisms of SMZ were elucidated. The Fe–Ce composite/H₂O₂ system exhibited extensive degradation capabilities towards SMZ as well as many other organic pollutants.

2. Experimental

2.1. Chemicals

The chemicals used in this work were as follows: sulfamethoxazole (SMZ, Aladdin), H₂O₂ solution (30% w/v, Aladdin), potassium dihydrogen phosphate (KH₂PO₄, Aladdin), 5,5-dimethyl-1-pyrroline N-oxide (DMPO, Sinopharm), nitroblue tetrazolium (NBT, Sinopharm), tert-butyl alcohol (TBA, Sinopharm), potassium ferricyanide (K₃[Fe(CN)₆], Sinopharm), cerous chloride (CeCl₃, Sinopharm), sublimed sulfur (S, Sinopharm), hydrazine hydrate (H₂N₄·H₂O, Sinopharm). All chemicals were used without further purification.

2.2. Synthesis of Fe–Ce composites

Fe–Ce-PBA was synthesized following the procedure outlined in our previous publication.¹⁹ A solution of Fe–Ce-PBA (0.1 g) was dispersed completely by sonication in a mixture of hydrazine hydrate (4 mL) and deionized water (6 mL). Subsequently, a specific amount of sulfur powder (the mass ratio of Fe–Ce-PBA to S powder: 3 : 1) was added to the dispersion and vigorously stirred for 2 h before being transferred to a Teflon-lined steel autoclave with a capacity of 50 mL. The autoclave was heated to 160 °C and maintained at that temperature for 5 h. The obtained Fe–Ce composites (designated as Fe–Ce–S–N) were collected, washed with deionized water and ethyl alcohol multiple times, and dried overnight in a vacuum at 60 °C. Fe–Ce–N and Fe–Ce–S–H were synthesized by the same process as Fe–Ce–S–N, with the difference that sulfur powder or hydrazine hydrate was not used, respectively (Scheme 1).

Scheme 1 The preparation processes of Fe–Ce–S–N.

2.3. Experiment procedures

In a typical Fenton-like experiment, 10 mg of the prepared catalyst was added to 50 mL of target pollutant solution (20 mg L⁻¹, pH = 5.0) and dispersed by ultrasound for 5 minutes. The sample solution was then stirred for an additional 30 minutes to establish adsorption–desorption equilibrium before introducing a specific amount of H₂O₂ into the reaction system. At certain time intervals, 0.8 mL of the sample solution was extracted and filtered through a needle filter with a pore size of 0.22 μm before being thoroughly mixed with methanol.

The SMZ concentration was determined using high-performance liquid chromatography (HPLC) equipped with a C₁₈ column and a UV-vis detector. The column temperature was maintained at 30 °C, and the injection volume was set to 10 μL. A mobile phase consisting of KH₂PO₄ solution (pH 3.0, 5 mmol L⁻¹)/acetonitrile (V/V = 60/40) was used at a flow rate of 1.0 mL min⁻¹, while detection wavelength was set at 270 nm with peak time around 5.5 min. Total organic carbon (TOC) analysis of the sample solution was performed using a TOC analyzer (Multi N/C 3100). Inductively coupled plasma atomic emission spectroscopy (ICP-AES) was used to measure the leaching metal ions in the sample solution. The concentration of H₂O₂ was determined spectrophotometrically using potassium titanium oxalate, the absorbance of the sample was measured at the maximum absorption wavelength of 386 nm. The fluorescence spectroscopy (FS, FP-6500) was used to semi-quantify the concentration of ·OH produced. Terephthalic acid reacted with ·OH to form 2-hydroxyterephthalic acid and showed fluorescence intensity at a wavelength of 426 nm, and the concentration of ·OH produced was proportional to the peak intensity.¹⁹

2.4. Characterization

The microstructure images of prepared samples were recorded on a Field emission scanning electron microscope (FESEM, SU8220, Hitachi, Japan) and high-resolution transmission electron microscopy (HRTEM, TF20, Joel 2100F). X-ray diffraction (XRD) patterns of the prepared catalysts were analyzed on a diffractometer (D8 Advance, Bruker) using Cu Kα radiation. The elemental composition and chemical oxidation state of the samples was determined by X-ray photoelectron spectroscopy (XPS, Escalab 250Xi, Thermo Fisher Scientific). The electron spin resonance (ESR) signals of ·OH and ·O₂⁻ spin trapped were performed on a spectrometer (ESR E500, Bruker) by spin-trap reagent of DMPO.

3. Results and discussion

3.1. Characterization

To assess the effect of hydrazine hydrate and sulfur powder on catalyst morphology evolution, the morphology and composition of Fe–Ce–S–N were further characterized by TEM and SEM. Compared to the typical cubic structure of Fe–Ce–S–H (Fig. S1a and b) and the sheet-like structure of Fe–Ce–N (Fig. S1c and d), Fe–Ce–S–N (Fig. S1e and f) showed a more fluffy and dispersed cluster-like structure, which was more favorable for the exposure of active sites and accelerated diffusion and mass transfer. Moreover, the lattice spacing of 0.330 nm and 0.191 nm corresponded to the (111), (220) crystallographic planes of FeS and CeS in Fe–Ce–S–N (Fig. 1b). EDS analysis showed that the surface of Fe–Ce–S–N contains five elements: Fe, Ce, N, O and S (Fig. 3c). The diffraction peaks in the XRD pattern of Fe–Ce–S–N (Fig. 1d) were assigned to the typical diffractions of FeS, CeS and FeCe₂O₄, which was consistent with EDS mapping.

Fig. 1 (a) TEM image, (b) HRTEM image of Fe–Ce–S–N and (c) EDS mapping of N, Fe, Ce, O and S; (d) XRD spectra of Fe–Ce–S–N, Fe–Ce–S–H and Fe–Ce–N.

Fig. 2a displayed the XPS spectra of Fe–Ce–N, Fe–Ce–S–H, and Fe–Ce–S–N. The Fe–Ce–S–N displayed obvious S 2p peaks compared with Fe–Ce–S–H (Fig. 2b), which illustrated the positive role of hydrazine hydrate in the introduction of S. Meanwhile, the S 2p high-resolution spectra of Fe–Ce–S–N showed three deconvolution peaks. The peaks at 166.8 eV and 163.8 eV were assigned to metal–S bonds, and the peak at 172.4 eV was ascribed to sulfur atoms of intermediate oxidation state.²⁰ As described in Fig. 2c, both Fe 2p^3/2 and Fe 2p^1/2 could be divided into double peaks, which corresponded to Fe(II) and Fe(III), respectively.²¹ The XPS spectrum of Ce 3d could be categorized into two groups (v and u), corresponding to the core-level spin–orbit splitting of Ce 3d^3/2 and Ce 3d^5/2, respectively (Fig. 2d). The peaks including v₁, v₃, v₄, u₁, u₃ and u₄ corresponded to the Ce(IV) state while the remaining two peaks referred to the Ce(III) state.²² The effective formation of metal–S bonds in Fe–Ce–S–N indicated that hydrazine hydrate not only created an alkaline environment to facilitate the decomposition of the precursor PBA, but also promoted as a reducing agent to reduce the S powder to S²⁻, thus facilitating the formation of metal sulfides.²³

Fig. 2 XPS spectra of Fe–Ce–S–N, Fe–Ce–S–H and Fe–Ce–N: (a) survey; (b) high-resolution of S 2p, (c) Fe 2p and (d) Ce 3d.

3.2. Catalytic activities assessment towards SMZ degradation

A series of experiments with different systems were conducted to evaluate the performance of Fe–Ce–S–N in degrading organic pollutants. As shown in Fig. 3a, the removal rate of SMZ by H₂O₂ or Fe–Ce–S–N system alone was minimal. The Fe–Ce–S–N/H₂O₂ system showed the best degradation efficiency for SMZ compared to other systems. Specifically, the Fe–Ce–N/H₂O₂ system showed a 34.5% decrease in the removal rate of SMZ within 30 min compared to the Fe–Ce-PBA/H₂O₂ system (Fig. 3a). The removal rate of SMZ by the Fe–Ce–N/H₂O₂ system was negative under the modification of hydrazine hydrate only. This might be due to the fact that the collapsed PBA framework obscured most of the active sites and hindered the mass transfer process. However, the removal rate of SMZ by Fe–Ce–S–H and Fe–Ce–S–N increased by 8.9% and 20.8%, respectively, after sulfidation, suggesting that S doping played a facilitating role in pollutant degradation. Among them, the k value of Fe–Ce–S–N was as high as 0.1863 min⁻¹, which was 2.1 and 8.1 times higher than that of Fe–Ce–S–H and Fe–Ce–N, respectively. (Fig. S2a†). It was noteworthy that the Fe and Ce leaching concentrations in the Fe–Ce–S–N structure after the reaction were 0.56 mg L⁻¹ and 0.17 mg L⁻¹, respectively, which were only 30.1% and 23.3% of those of Fe–Ce–S–H and 37.6% and 31.5% of those of the precursor PBA (Table S1†). This indicated that the PBA-derived catalysts had good stability under the combined action of hydrazine hydrate and S powder. To further verify the superiority of the catalytic performance of Fe–Ce–S–N, the performance of various metal-containing catalysts/H₂O₂ systems for the removal of sulfonamide pollutants was compared.^24–31 The catalytic performance of Fe–Ce–S–N in the presence of moderate H₂O₂ was favorable.

Fig. 3 (a) SMZ degradation curves in various reaction systems; (b) Comparison of the removal rate of sulfonamide pollutants with reported Fenton catalysts;^24–31 Fe–Ce–S–N/H₂O₂ system: (c) removal rate of SMZ under different initial pH, (d) removal rate of different pollutants, (e) TOC removal in 2 h and (f) cycling performances. Reaction conditions: [pollutants] = 20 mg L⁻¹, [catalyst] = 0.2 g L⁻¹, [H₂O₂] = 15 mM, pH = 5.0, T = 25 °C.

As displayed in Fig. 3c, efficient degradation of SMZ (>95% degradation within 30 min) was achieved at initial pH in the range of 3.11–8.95. When the pH was increased to 11.08, the degradation effect of SMZ was significantly inhibited, and the removal rate was only 12.2% in 30 min, which was mainly due to the formation of iron hydroxide precipitate and the ineffective decomposition of H₂O₂ due to the alkaline condition.³² Besides, the leaching concentration of Fe and Ce in the measured pH conditions was within the EU permitted emission standard (less than 2.0 mg L⁻¹) (Fig. S2b†). Excess H₂O₂ in Fe–Ce–S–N/H₂O₂ system would quench the active radicals (eqn (1) and (2)) and competed with SMZ for metal active centers on the surface of Fe–Ce–S–N, and thus hindering the catalytic degradation process (Fig. S3†).

·OH + H₂O₂ → ·OOH + H₂O (1)

·OHH + ·OH → O₂ + H₂O (2)

Excess catalysts could also have detrimental effects on the reaction. Therefore, considering the catalytic effect and the cost of chemicals, the optimal experimental conditions were determined to be pH = 5, 0.2 g L⁻¹ Fe–Ce–S–N and 15 mM H₂O₂.

To verify the applicability of Fe–Ce–S–N/H₂O₂ system for different organic pollutants, ciprofloxacin (CIP), bisphenol A (BPA) and methyl orange (MO) were selected as the model pollutants to evaluate their degradation performance. Fe–Ce–S–N/H₂O₂ system could completely remove different types of organic pollutants within 30–45 min (Fig. 3d). Meanwhile, the TOC removal rate of these pollutants ranged from 43.7% to 58.3% after 2 h (Fig. 3e), indicating the superior potential application of Fe–Ce–S–N/H₂O₂ system.

The reusability and stability of catalysts are the key factors for their practical application. Herein, in order to verify the reusability of Fe–Ce–S–N, the cycling experiment was performed. The degradation efficiency of SMZ could still reach to 97% after four cycles (Fig. 3f), suggesting the excellent reusability and stability of Fe–Ce–S–N. The little loss of catalyst and the adsorption of intermediates on the active site may be the possible reasons for the slight decrease of degradation efficiency. Besides, as shown in Table S2,† after the fourth repetition, the Fe and Ce release was only 0.19 and 0.02 mg L⁻¹. Furthermore, the positions of the characteristic FT-IR peaks of the reacted materials did not change significantly and retained the characteristic functional structure of the sulfide, illustrating the catalytic stability of Fe–Ce–S–N/H₂O₂ system (Fig. S5†). Moreover, Fe–Ce–S–N can maintain high catalytic activity in complex aqueous environments where lower concentrations of anions and organic matter coexist (Fig. S6†), and has the potential to be applied to the removal of organic pollutants in real water.

3.3. Mechanisms of Fe–Ce–S–N/H₂O₂ system

To clarify the mechanism of H₂O₂ activation and SMZ degradation in the Fe–Ce–S–N/H₂O₂ system, the primary active species were confirmed by quenching experiment and ESR characterization. The degradation efficiency of SMZ decreased by 81.5% in the presence of excess TBA within 30 min, while that decreased by 65.1% in the presence of excess NBT (Fig. 4a). It meant that ·OH and ·O₂⁻ were the main active species in Fe–Ce–S–N/H₂O₂ system, which was consistent with the ESR characterization (Fig. 4c and d).³³ The decomposition rate of H₂O₂ by different catalysts in Fenton-like reaction process was investigated. As shown in Fig. 4b, 60.5% of H₂O₂ decomposition was achieved after 60 min reaction in the Fe–Ce–S–N/H₂O₂ system, while only 28.7% of H₂O₂ was decomposed with the addition of Fe–Ce–N, and the decomposition rate could reach to 49.8% with additive Fe–Ce–S–H. Moreover, FS spectra further verified the changes in ·OH production with time, demonstrating the high efficiency of Fe–Ce–S–N in activating H₂O₂ to ·OH.

Fig. 4 (a) SMZ degradation in the presence of various active species scavengers; (b) Evolution of H₂O₂ concentration and ·OH production (inset) in Fe–Ce–S–N, Fe–Ce–S–H and Fe–Ce–N based Fenton-like systems; EPR spectra of (c) DMPO–·OH and (d) DMPO–·O₂⁻ in Fe–Ce–S–N/H₂O₂ systems.

The ratio of Fe and Ce in different valence states before and after use was analyzed by XPS and depicted in Fig. S7 and Table S3.† The increase in the proportion of Fe(II) and the decrease in the ratio of Ce(III) after the reaction indicating that Ce(III) acted as electron donor in Fe–Ce–S–N/H₂O₂ system, and electrons of Ce(III) would migrate to Fe(III) through Fe–S bond. Moreover, due to the participation of S, the conversion of Fe(III) to Fe(II) was further promoted, which accelerated the Fe cycle. The increased cycling efficiency of Fe(III) to Fe(II) favored the degradation process (Fig. 5). The Fe–Ce–S–N/H₂O₂ system mainly included two mechanisms: (i) Fe–Ce–S–N catalyst had multiple active sites, and Fe(II), Fe(III), Ce(IV) and Ce(III) on the catalyst surface could activate H₂O₂ to produce active oxygen species;^34,35 (ii) The redox potential of Fe(III)/Fe(II) (0.7 V/NHE) is higher than that of S/S²⁻ (−0.48 V/NHE), thus, S²⁻ could act as an electron donor to promote Fe(III)/Fe(II) cycle.^36–38 Herein, the effective cycle and regeneration of active sites over the Fe–Ce–S–N resulted in a stable and efficient Fenton-like activities towards SMZ degradation.

Fig. 5 The possible mechanism in the in Fe–Ce–S–N/H₂O₂ system.

3.4. Degradation products, pathways and toxicity analysis

The possible degradation products and degradation pathways for SMZ were depicted in Fig. 6. The hydroxylation, opening of isoxazole ring and cleavage of SMZ were confirmed as the primary degradation mechanisms. In pathway I, the methyl group in the isoxazole ring of SMZ was replaced by the hydroxyl group and formed the hydroxylated intermediate (P1, m/z = 256) by the attack of ·OH, while the pathway II was due to the dehydrogenation of amino groups in the benzene ring of SMZ (P15, m/z=284). The intermediate (P2, m/z =246) was produced by the opening of isoxazole ring, and its further cleavage could form the intermediates 3 (P3, m/z = 216), 4 (P4, m/z = 198) and 5 (P5, m/z = 158). In pathway IV, the benzene ring of SMZ was attacked by ·OH and produced the hydroxylated intermediate (P6, m/z = 256). The intermediates 7 (P7, m/z = 190) and 8 (P8, m/z = 99) were formed through the rupture of S–N bond in the intermediate 6. Meanwhile, the intermediate 8 could generate the intermediates 9 (P9, m/z = 115) and 10 (P10, m/z = 111) through isoxazole ring hydroxylation and amino group dehydrogenation, respectively. And the intermediate 9 (P9, m/z = 115) could produce the intermediates 11 (P11, m/z = 133), 12 (P12, m/z = 117), 13 (P13, m/z = 118) and 14 (P14, m/z = 102) by a series of reactions including hydroxylation, dehydrogenation, ring opening and bond cleavage.

Fig. 6 Possible degradation pathways of SMZ in Fe–Ce–S–N/H₂O₂ system.

Investigation into the toxicity of degradation products was essential for evaluating the environmental risk associated with degradation reactions. Therefore, the biological toxicity of SMZ and its possible intermediate products was estimated by the toxicity assessment software (ECOSAR) based on quantitative structure–activity relationship (QSAR). According to Table S4,† the classification of acute and chronic toxicities can be defined as very toxic, toxic, harmful, and not harmful. The predicted results were presented in Table S5† and Fig. 7. In terms of acute toxicity, most of the intermediate products were referred to as “Not harmful” compounds for the three target species. Meanwhile, only a few intermediate products exhibited obvious higher chronic toxicities than SMZ, such as the predicted ChV to fish of P4, the ChV to Daphnid of P8, and the ChV to Green Alage of P14. Herein, the degradation reaction of this study was of relative low environmental risk.

Fig. 7 The predicted toxicity values of SMZ and its products.

4. Conclusions

In summary, we successfully synthesized S-modified clustered Fe–Ce–S–N composites by a simple hydrothermal reaction. The obtained Fe–Ce–S–N catalysts, prepared under the optimal conditions, exhibit a unique composition with mixed phases of FeS, CeS and FeCe₂O₄ that were favorable for the improvement of Fenton-like activities. Benefitting from these merits, the Fe–Ce–S–N materials demonstrate excellent Fenton-like activities with a reaction rate constant as high as 0.1863 min⁻¹, and also exhibit good long-term stability and high efficiency for various organic pollutants, indicating their great potential as promising candidates for practical wastewater treatment. The findings of this study can provide valuable insights into designing highly efficient Fenton-like catalysts and understanding the mechanism of H₂O₂ activation and contaminant degradation.

Author contributions

Jinli Qiu: conceptualization, investigation, writing–original draft. Huifang Chan: data curation, visualization, writing–review & editing. Wenting Zheng: experiment, methodology. Yao Feng: writing–review & editing. Fuqiang Liu: supervision, conceptualization, validation, writing – review & editing.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

The authors gratefully acknowledge the support of National Natural Science Foundation of China (No. 51522805), The Natural Science Research Project of Universities in Jiangsu Province of China (No. 22KJB610005), Natural Science Foundation of Nanjing Normal University (No. 184080H202B354) and The National Key Research and Development Program of China (No. 2023YFE0100900). We would like to thank the researchers in the Shiyanjia Lab (https://www.shiyanjia.com/) for their helping with XPS analysis.

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Footnotes

† Electronic supplementary information (ESI) available. See DOI: https://doi.org/10.1039/d3en00788j

‡ Jinli Qiu and Huifang Chan are co-first authors.

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