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Article

Facile Preparation of Fe3O4@SiO2 Derived from Iron-Rich Sludge as Magnetic Catalyst for the Degradation of Organic Contaminants by Peroxymonosulfate Activation

by
Zhiwei Wang
1,
Shijun Zhu
1,*,
Jing Deng
1,
Haojie Li
1,
Liang Wang
2,
Haojin Luo
3,
Zehe Tang
3 and
Xueyan Li
4
1
College of Civil Engineering, Zhejiang University of Technology, Hangzhou 310023, China
2
The Architectural Design and Research Institute of Zhejiang University Co., Ltd., Hangzhou 310028, China
3
Wenzhou Design Group Co., Ltd., Wenzhou 325000, China
4
School of Environmental Science and Engineering, Suzhou University of Science and Technology, Suzhou 215009, China
*
Author to whom correspondence should be addressed.
Sustainability 2022, 14(24), 16419; https://doi.org/10.3390/su142416419
Submission received: 18 November 2022 / Revised: 2 December 2022 / Accepted: 7 December 2022 / Published: 8 December 2022
(This article belongs to the Special Issue Application of Oxidation Technology for Water Treatment)
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Abstract

:
Iron-rich sludge, generated during flocculation/sedimentation processes by using Fe-based coagulant in drinking water treatment plants, could be used as a precursor to prepare an effective peroxymonosulfate (PMS) activator (Fe3O4@SiO2) for the ciprofloxacin (CIP) degradation via facile hydrothermal treatment. The catalytic performances of raw iron-rich sludge and Fe3O4@SiO2 were evaluated. The removal rate of CIP in Fe3O4@SiO2/PMS system increased from 44.7% to 82.8% within 60 min compared with the raw iron-rich sludge. The effects of PMS, catalyst loadings, temperature, and initial pH on the CIP degradation were examined, demonstrating that acidic conditions and higher temperatures were beneficial for CIP degradation. Both sulfate radicals (SO4•−) and hydroxyl radicals (•OH) contributed to the CIP degradation, and SO4•− was predominated in the Fe3O4@SiO2/PMS system, which was confirmed by the result of electron paramagnetic resonance (EPR) analysis and radical quenching tests. The mechanisms of the PMS activation process by Fe3O4@SiO2 were elucidated, and the influencing factors were among which the role of the iron mineral phase was emphatically explored. This study provides a facile method to convert the recycled waste iron-rich sludge to magnetic heterogeneous catalysts for CIP degradation with PMS activation.

1. Introduction

Coagulation/flocculation-sedimentation processes play a significant role in the removal of colloidal solids, natural organic matter (NOM), and pathogens in drinking water purification [1]. The use of coagulants results in the generation of large quantities of drinking water treatment sludge, which requires handling and, ultimately, disposal. Approximately 10,000 tons of DWTS are produced each day worldwide [2], and it exceeds 130,000 and 260 million wet tons in the United Kingdom and China per year, respectively [3]. The majority of DWTS is still disposed to landfills and discharged into the sewers or water environments without any treatment [4,5], which poses a potential risk to aquatic environments and triggers long-term secondary pollution in soil [6]. Therefore, more feasible sludge reduction and re-use strategies of DWTS should be considered.
Iron salts as common coagulants are also widely used in drinking water treatment plants. Thus, the DWTS are mesoporous multiphase amorphous materials rich in iron. Utilizing iron-rich sludge to synthesize new materials for economic element adsorption and recovery has become a new research direction. The iron-rich sludge has a well-developed pore size, various functional groups, and a large specific surface area, which has potential as an adsorbent for heavy metals [7], organic pollutants [8], and phosphate [9]. In addition, this waste iron sludge has superior catalytic moieties, porous structure, hydroxyl functional groups, and other active sites on the surface. Considering the high iron concentration of iron-rich sludge (where iron salts are used as coagulants), it has the potential to be beneficially reused as feedstock or a precursor to develop a cost-effective environmental functional material, due to: (i) the high concentration of oxides of transition metals, sufficient surface areas, and abundant functional groups for a number of catalytic applications [10,11]; and (ii) it does not require high energy input and the massive addition of fresh chemicals, which may help to reduce second pollution [12].
Advanced oxidation processes (AOPs) based on reactive oxygen species (ROS) have received increasing scientific attention in the degradation of refractory organic pollutants [13]. In persulfate (PS)-based AOPs, peroxymonosulfate (PMS) and peroxydisulfate (PDS) could be activated to generate various ROS. In comparison to hydroxyl radicals (•OH), the sulfate radical (SO4•−) has a higher standard redox potential (SO4•−: 2.5–3.1 V, •OH:1.8–2.7 V), a wide solution pH range (2–8), and a longer half-life time (30–40 μs) [14,15]. SO4•− could be effectively generated through the activation of persulfate (PS) with heating [16], ultrasound [17], ultraviolet radiation [18], transition metal ions [19], and carbon materials [20]. There are several disadvantages of homogeneous iron-activated PMS systems, including strict pH restriction, large input of iron ion dosage, and the introduction of anions related to iron salts like SO42− or Cl−. Considerable efforts have been made to develop various transition metal-based heterogeneous catalysts for PS activation. Among them, Fe-mediated activation of PS has been receiving a lot of interest ascribed to its low toxicity, abundance, and superior catalytic activity. For example, numerous researches have focused on the synthesis of heterogeneous Fe-based catalysts, such as Fe3O4, due to the fact that non-toxic and abundant Fe(II) could induce the decomposition of PMS to produce SO4•− and •OH. Moreover, Fe3O4 with magnetic property could be easily separated from aqueous solution in the presence of an external magnetic field [21]. Biochar derived by pyrolyzing biomass under oxygen-limited conditions can behave as an effective carrier for transition metals, and this type of biochar-loaded iron oxides composites exhibit higher catalytic activity [22]. However, the sewage sludge-derived biochar has to be developed by external additional Fe sources and high pyrolysis temperature (>800 °C), which inhibits the practical application. Especially, iron-containing wastes like DWTS, as PMS activators, have also attracted great attention. The high Fe concentration makes it possible for iron-rich sludge to be an excellent catalyst. Furthermore, the metal nanoparticles can be settled with the hydroxyl groups on the sludge surface, which would promote PMS activation [23]. Iron-rich sludge contains iron hydroxide (Fe(OH)3 or FeOOH), silica, and organic carbon. Hence, it has a great potential to be beneficially recycled as a sufficient resource to prepare a cheap and sustainable material with low chemical and energy consumption for the decontamination of pollutants. Hydrothermal treatment is considered a promising approach for the preparation of catalysts due to its advantages of moderate reaction conditions, low energy input, simple process, and convenient operation [24].
Powdered solid iron-rich sludge might act as the precursor and support to synthesize a green and economic PMS catalyst without high energy input and a massive addition of transition metals. Therefore, exploring its additional value (as an iron source and catalyst support) while dealing with this iron-rich sludge may be a process of waste resource utilization. Herein, in this work, a facile hydrothermal treatment was used to convert partial Fe(III) species in raw iron-rich sludge into Fe(II) through this reduction process. The as-prepared magnetic nanocomposite (Fe3O4@SiO2), possibly with superior catalytic activity, a porous structure, and large specific surface area, behaves as a PMS activator for the oxidation of refractory organic pollutants in wastewater. Ciprofloxacin (CIP), as a typical fluoroquinolone antibiotic, was selected as the target organic compound due to the frequent detection in various aquatic environments and inefficient removal by traditional treatment techniques [25]. Herein, the catalytic performance of PMS activation by the raw iron-rich sludge and Fe3O4@SiO2 were evaluated. Meanwhile, the crystalline structure, morphology, and other physicochemical properties of iron-rich sludge and Fe3O4@SiO2 were characterized. The effects of PMS concentration, catalyst dosages, reaction temperature, and initial pH on CIP degradation efficiency were also examined. Electron paramagnetic resonance (EPR) analysis and radical quenching tests were used to identify the predominant reactive species produced in the Fe3O4@SiO2/PMS system. The potential catalytic mechanisms in the Fe3O4@SiO2/PMS system were explored based on the analysis of EPR and radical quenching tests. The present study intends to provide a sustainable strategy for the reuse of iron-rich sludge to develop a high-efficiency and low-consumption catalyst for the removal of organic contaminants via PMS activation with the aim of “treating wastewater by waste sludge”.

2. Materials and Methods

2.1. Reagents and Chemicals

Ciprofloxacin (CIP ≥ 99.9%), PMS (Oxone, KHSO5·0.5 KHSO4·0.5 K2SO4), and PDS (K2S2O8) were obtained from Sigma-Aldrich (Shanghai, China). Ethylene glycol (EG), Ethanol (EtOH), and tert-butyl alcohol (TBA) were purchased from Aladdin Chemical Reagent Co., Ltd. (Shanghai, China). HPLC grade acetonitrile and acetic acid were received from Merck and Thermo Fisher Scientific, respectively. Other reagents were purchased from Sinopharm Chemical Reagents Co., Ltd., Shanghai, China. All solutions were freshly prepared with deionized water (>18.2 MΩ·cm) from a Millipore ultrapure purification system.

2.2. Preparation of Catalysts

The raw iron-rich sludge was collected from the sedimentation tank of a practical drinking water treatment plant by using commercial FeCl3 as the coagulant. The precipitated iron-rich sludge was re-aggregated with polyacrylamide (PAM), then dewatered and concentrated through centrifuges and a frame filter press, subsequently naturally air-dried outside for several days and dried in an air oven at 105 °C for 24 h to remove the moisture. The dry iron-rich sludge was grinded and sieved to 200 mesh, then stored in a dry pan and used as the material for subsequent modification. The content of Fe in the ferric sludge was found to be 188.6 ± 3.50 mg/g. The content of compositions was measured by ICP-AES and XPS analysis. The main characteristics of the iron-rich sludge are shown in Figure S1 and Table S1.
The Fe3O4@SiO2 was synthesized by using a hydrothermal treatment process as follows: 0.7 g dry iron-rich sludge was mixed with 35 mL ethylene glycol with 3.6 g sodium acetate (NaAc) added as ligand, and thoroughly mixed by mechanical stirring and ultrasonic. Secondly, the mixture was transferred to 50 mL Teflon-sealed autoclaves after being shaken for 60 min and then heated to 200 °C at a rate of 4 °C/min for 10 h. After the autoclave cooling to room temperature, the obtained products were collected and washed with ultrapure water and ethanol several times, and, finally, dried in a vacuum at 60 °C for 12 h.

2.3. Characterization Techniques

The surface morphology of raw iron-rich sludge and Fe3O4@SiO2 were identified by scanning electron microscopy (FE-SEM; Zeiss SUPRA 55 SAPPHIRE, Berlin, Germany) equipped with an Energy Dispersive Spectrometer (EDS) and a High-Resolution Transmission Electron Microscope (HR-TEM; JEM-2100F, JEOL Ltd., Tokyo, Japan). The crystal structure of the catalysts was studied by X-ray powder diffraction (XRD) with Cu Ka radiation (Bruker D8, λ = 0.15418 nm). Fourier transform infrared spectroscopy (FTIR) was used to analyze the chemical functional groups and chemical bounds of catalysts in the 4000–400 cm−1 region (Perkin-Elmer Spectrum One B spectrometer, Rheinstetten, Germany). An automatic microspore physisorption analyzer (Tristar 3020, Micromeritics, Norcross, GA, USA) was conducted to measure the specific surface areas of the samples. The magnetic properties of raw iron-rich sludge and Fe3O4@SiO2 were measured by using a vibrating sample magnetometer (VSM, Lake Shore 7304, Cedar Lake, IN, USA).

2.4. Experimental Procedures

The experiments were performed in 250 mL glass beakers containing 100 mL CIP solutions (10 mg/L) with magnetic stirring at a steady speed of 400 rpm at a temperature of 25 ± 0.2 °C. The initial pH of batch experiments was adjusted with 0.1 M HClO4 and NaOH solution. The reaction was initiated by quickly adding catalysts (0.2 g/L) and persulfate (1.6 mM PMS or 2.0 mM PDS solution). Then, 1.0 mL of aqueous sample was withdrawn at fixed time intervals (0, 1, 2, 5, 10, 15, 20, 30, 45, and 60 min) and immediately stored in a 2.0 mL centrifuge tube existing 0.5 mL NaNO2 solution (1 g/L) to quench the reaction. Then, 1 mL water samples were withdrawn and filtered through 0.22 µm polytetrafluoroethylene filters prior to analysis. All experiments were conducted in duplicate.

2.5. Analytical Methods

The residual concentrations of CIP were analyzed by a High-Performance Liquid Chromatography (HPLC, Agilent 1220, Santa Clara, CA, USA) system equipped with a Waters BEH C18 column (100 mm × 2.1 mm, 1.7 µm) at the flow rate of 1 mL/min and 298 K (with a mobile phase of acetonitrile/water (20/80, v/v) and a detecting wavelength at 278 nm). The free radicals generated in the reaction system were detected by electron paramagnetic resonance (EPR) spectroscopy (Bruker A200 spectrometer, Rheinstetten, Germany) using 10 mM 5,5-dimethyl-1-pyrroline N-oxide (DMPO) as a spin-trapping agent (JES-FA200, JEOL, Tokyo, Japan).

3. Results and Discussion

3.1. Characterization of Raw Iron-Rich Sludge and Fe3O4@SiO2

Before the solvothermal treatment, iron-rich sludge was pretreated and analyzed. The content of compositions was measured by ICP-AES and XPS analysis. The content of Fe in the ferric sludge was found to be 188.6 ± 3.50 mg/g. The main characteristics of the iron-rich sludge are shown in Table S1. XPS was employed to investigate the percentage of Fe in different valence states in the catalyst. As displayed in Figure S1, the peaks situated at around 710.0 and 723.6 eV corresponded to Fe(II) species (19.1 atom%), while the binding energies of 712.1 and 725.7 eV were assigned to Fe(III) species (80.9 atom%), suggesting the coexistence of Fe(II)/Fe(III) on the surface of the catalyst. Thus, the Fe(II) and Fe(III) concentration could be calculated as 36.0 ± 0.67 mg/g and 152.6 ± 2.83 mg/g, respectively.
The crystal structure of relevant samples was characterized by XRD. As presented in Figure 1a, the characteristic diffraction peaks at 2θ = 20.9° and 26.7° indicated the existence of silica (SiO2) during the solvothermal reaction. The SiO2 in magnetic catalysts derived from iron-rich sludge has been extensively investigated as the support for the transition metal-based catalyst (i.e., Fe3O4, Fe2O3) due to the Fe-O-Si interactions, which could improve the stability and reduce iron ion leaching [26]. However, the reduction of SiO2 peak intensity might be attributed to the etching effect of weak alkaline conditions in the presence of NaAc. In addition, the characteristic diffraction peaks at 2θ = 30.3°, 35.7°, 43.3°, 57.3°, and 63.0° corresponded to (220), (311), (222), (511), and (214) planes of Fe3O4 (JCPDS 76-1849) [27].
The functional groups of Fe3O4@SiO2 and iron-rich sludge samples were identified by FTIR (Figure 1b). The broad peak at 3380 cm−1 reflects the stretching vibration of O-H bonds [28], while the peak that appeared at 1630 cm−1 and 1410 cm−1 were attributed to the C=O and O=C-O bonds [27]. Furthermore, the bonds at 466 cm−1 were assigned to the Si-O-Fe bonds, and the C-O bonds were observed at 1010 cm−1 [29]. Compared with raw iron-rich sludge, the vibrational peak of the -OH bond in Fe3O4@SiO2 became weakened due to the evaporation of surface molecular H2O during the solvothermal treatment process. The decrease of -CH3 peak intensity could be explained by the decomposition of organic matters and algae cells in iron-rich sludge. The decrease in O=C-O and C-O stretching vibration was related to the decarboxylation after hydrothermal reaction. Furthermore, stretching vibrations of Si-O-Fe and Fe-O functional groups represented the formation of Fe3O4 crystals loading on the SiO2.
The N2 adsorption/desorption isotherms and specific surface area of the Fe3O4@SiO2 and iron-rich sludge were measured. As shown in Figure 1c, the N2 adsorption/desorption isotherm of Fe3O4@SiO2 presented a typical type IV curve with an obvious H3 hysteresis loop at p/p0 range from 0.5 to 0.9 [30], which was obviously different from that of iron-rich sludge. This result illustrated that the catalyst Fe3O4@SiO2 has a typical mesoporous structure with an average pore size of 5.76 nm. Based on the BET models, the surface area (SBET) of Fe3O4@SiO2 (50.352 m2/g) was about 1.3 times higher than that of raw iron-rich sludge (21.486 m2/g) (Figure 1d). The pore volume of Fe3O4@SiO2 increased from 0.059 cm3/g to 0.103 cm3/g, which indicated that the hydrothermal treatment promoted the micro-porous structural properties of Fe3O4@SiO2. As shown in Table 1, the increase of SBET was beneficial to the exposure of active sites and increased the contact probability between the oxidants and organic molecules, which could improve the catalytic performance further [31].
The surface microstructure and morphology of raw iron-rich sludge and Fe3O4@SiO2 were analyzed by SEM and TEM. As shown in Figure 2a, iron-rich sludge displayed a dense structure and irregular morphology. The morphology of biological cells from algal cells in the raw water was observed. The TEM image (Figure 2b) also confirmed the interior structure of the iron-rich sludge was amorphous and uneven, which may indicate the existence of non-crystalline hematite and other inorganic minerals, such as feldspar, calcite, kaolin, and silicate. The details of chemical element components of iron-rich sludge are listed in Table S2 based on EDS mapping analysis (Figure S2), it is suggested that C, O, N, Fe, and Si were the main elements accounting for about 80% in raw sludge. Particularly, the tiny crystal sphere with a size ranging from 50 to 100 nm was uniformly distributed on the catalyst surface after hydrothermal treatment, which could confirm the formation of Fe3O4 (Figure 2c). The TEM image (Figure 1d) of Fe3O4@SiO2 also exhibited a nano-spherical structure with an average diameter of 80 nm without obvious aggregation.
TGA-DSC profiles of iron-rich sludge in air atmosphere are shown in Figure 3a. The weight loss ultimately reached about 30% with the increase of pyrolysis temperature. The mass loss occurred through three stages at various temperatures: (i) the weight loss efficiency in the first stage was about 7.5% with increasing room temperature to 220 °C due to the release of bound water and decomposition of hydrocarbons (CH4) and hydroxyl phase (-OH); (ii) the primary mass loss occurred at the second stage was 18.5% when the temperature rose from 220 °C to 400 °C, which was ascribed to the carbonization of natural organics and release of CO2 and H2O; (iii) around 4.2% of weight loss further occurred during the third stage, suggesting the decomposition of calcium-magnesium carbonate. Eventually, the Fe species transferred to Fe2O3 in the presence of oxygen. The magnetization curves of Fe3O4@SiO2 were analyzed by VSM (Figure 3b). Compared with the iron-rich sludge, the catalyst after hydrothermal treatment exhibited superior magnetic properties with a saturation magnetization value of 9.5 emu/g. These magnetic parameters suggest that the Fe3O4@SiO2 could be effectively separated from the treated water by an external magnetic field.

3.2. Evaluation of CIP Removal by Raw Iron-Rich Sludge

3.2.1. Adsorption Capacity of Iron-Rich Sludge

The CIP removal efficiency via physical adsorption by raw iron-rich sludge was evaluated. As shown in Figure 4a, less than 10% of CIP was removed within 60 min when the dosage of iron-rich sludge was at a range of 0.2~0.5 g/L. However, the adsorption efficiency of CIP improved from 10% to more than 80% with increase of iron-rich sludge dosages from 0.2 to 5.0 g/L. According to Equation (S1) (Text S1), the adsorption capacity for CIP under different dosages of iron-rich sludge has been calculated and is presented in Figure 4b. The saturated adsorption capacity for CIP could be maintained when the iron-rich sludge dosage exceeded 0.5 g/L.
The mechanisms of CIP adsorption were primarily dependent on the chelation of Fe on the surface of iron-rich sludge with the carboxyl group (-COOH) in CIP molecules, as well as the hydrogen bonding between -OH and O=C (Figure 5). However, a large dosage of raw sludge would lead to effluent quality deterioration and secondary pollution. Furthermore, CIP molecules were only transferred from the liquid phase to the solid phase via physical adsorption rather than being completely degraded. Therefore, the adsorption of CIP by iron-rich sludge alone is not a feasible method in terms of practical application.

3.2.2. Persulfate Activation by Iron-Rich Sludge

The catalytic efficiency utilizing PMS and PDS as oxidants of iron-rich sludge was investigated (Figure 6). Only 21.4% and 24.3% of CIP (10 mg/L) were removed in 60 min by PMS and PDS oxidation alone, respectively. The simultaneous presence of iron-rich sludge (200 mg/L) and PMS (1.6 mM) sightly stimulated the degradation, with 44.7% of CIP yielded within 60 min, while 51.7% of CIP was removed in the iron-rich sludge/PDS system. The inefficiency of catalytic activities by iron-rich sludge could be explained in that the main form of iron in iron-rich sludge is Fe(III), which is converted into Fe(II) first via a one-electron reduction process that occurs on the surface of the catalyst before two persulfates (PMS and PDS) are activated (Equations (1) and (2)) [12]. Then, the resulting Fe(II) directly activates persulfates to produce free radicals for organic degradation (Equation (3)). However, the slow conversion rate of Fe(III) to Fe(II) and persulfate consumed during the reaction both lead to the lower catalytic efficiency of Fe(III) oxide than Fe(II) oxide.
Fe(III)-OH + S2O82− → Fe(II)-OH + S2O8•−
Fe(III)-O + S2O82− → Fe(II)-O− + S2O8•−
Fe(III) + HSO5− → Fe(II) + SO5•− + H+
In addition, the influence of iron-rich sludge and PMS dosages on CIP removal was investigated. As shown in Figure 7a, the removal efficiency of CIP improved from 61.4% to 94.6% within 60 min with the increase of iron-rich sludge dosage from 0.2 g/L to 2 g/L. Additionally, the degradation of CIP was enhanced from 70.2% to 97.6% within 60 min as the PMS concentration increased from 0.75 g/L to 2 g/L (Figure 7b). When the dosages of PMS were 1 g/L and 2 g/L, the degradation trend of CIP in 15 min exhibited the same tendency due to the limited surface-active sites of catalyst. This phenomenon demonstrated that high CIP removal efficiency could be achieved by increasing the dosage of iron-rich sludge and PMS owing to more exposed active sites for activation of PMS and the generation of free radicals. However, large quantities of dosage will raise the cost of water treatment investments, making it difficult to be practically applied in water treatments. Therefore, it seems necessary for iron-rich sludge to be further modified to improve the catalytic activity for organic degradation.

3.3. Catalytic Performance of Fe3O4@SiO2

CIP degradation utilizing Fe3O4@SiO2 as a catalyst in different systems has been evaluated. As in Figure 8a, Fe3O4@SiO2 alone could eliminate 19.4% of CIP in 60 min, implying the inferior absorption capacity of Fe3O4@SiO2. Only 21.4% of CIP were removed by PMS alone due to its low redox potential (E0 = 1.82 V) [32]. Compared with raw iron-rich sludge, Fe3O4@SiO2 has shown prior catalytic ability towards PMS, and 82.8% of CIP was degraded within 60 min in the Fe3O4@SiO2/PMS system. The result indicated that the catalytic activity of Fe3O4@SiO2 could be improved due to the partial transformation from Fe(III) to Fe(II) after hydrothermal reduction reaction. Furthermore, the larger specific surface area of Fe3O4@SiO2 also contributed to the increase of active sites for PMS activation.
Fe3O4@SiO2 was employed for comparison of the activation performance between two persulfates (i.e., PMS and PDS). As shown in Figure 8b, the CIP degradation efficiency was found to be 62.9% within 60 min in the Fe3O4@SiO2/PDS system, which was lower than that by the PMS activation system. The result illustrated that Fe3O4@SiO2 was highly effective for PMS activation as PMS possesses a higher oxidation potential and asymmetric molecule structure. The longer superoxide O-O bond (lO-O = 1.326 Å) and the asymmetric structure (HO-O-SO3−) has been proved easier for PMS activation to produce free radicals than PDS, which has a symmetric structure (−3OS-O-O-SO3−) and a shorter O-O bond (lO-O = 1.322 Å) [33].
The effects of Fe3O4@SiO2 and PMS dosages on the CIP degradation were explored. Figure 9a showed that the CIP removal efficiency was improved from 65.7% to 82.8%, 97.6%, and 98.6% within 60 min with an increase in Fe3O4@SiO2 dosage from 0.1 g/L to 0.2 g/L, 0.5 g/L, and 1 g/L. When the catalyst dosage was less than 0.2 g/L, PMS could not be adequately activated to generate sufficient free radicals due to the limited concentration of Fe(II) from Fe3O4@SiO2 in this system. CIP removal efficiency was enhanced from 56.5% to 82.8% and 86.3% within 60 min with the PMS concentration increasing from 0.25 g/L to 0.5 g/L and 0.75 g/L (Figure 9b). Additionally, the degradation efficiency of CIP improved to 92.5% in 30 min as the PMS dosage further increased to 1 g/L. The lack of PMS made it hard to generate free radicals when the dosage was below 0.75 g/L. The results determined that elevating the dosages of Fe3O4@SiO2 and PMS could increase more exposed active sites for the PMS activation and the collision frequency between the Fe3O4@SiO2 and PMS, which could result in more reactive free radicals generation and enhancement on organic compounds degradation efficiency. However, the overdose of PMS and catalysts could only slightly improve the CIP degradation efficiency, likely ascribed to the excessive radical self-scavenging or ineffective PMS decomposition at high catalyst dosage [34]. Accordingly, 200 mg/L catalyst and 0.5 g/L PMS were used in the subsequent experiments.

3.4. Effects of Initial pH and Temperature

One of the most critical factors influencing the catalytic ability is the initial pH of the reaction solution [35]. As depicted in Figure 10, the CIP could be entirely removed at pH 3 and 5 in the Fe3O4@SiO2/PMS system. However, the process seemed to have slowed at pH 9, with just 60% of the CIP being eliminated in 60 min. These results revealed that the acidic condition was more favorable to the CIP degradation in the Fe3O4@SiO2/PMS system, which could be explained from the following: (1) The acidic condition is beneficial for the corrosion of iron oxide and acceleration of Fe2+ release into the reaction solution [36], while the rise in pH results in the formation of hydroxyl groups or iron complexes (i.e., Fe(OH)x and FeOOH) on the catalyst surface, which may cover and passivate the active sites of catalyst [37]. (2) According to Equations (4) and (5), more SO4•− will be transformed to •OH with an increase of pH value. Therefore, the alkaline conditions partially inhibited the total oxidation capacity of the Fe3O4@SiO2/PMS system for CIP removal due to the lower redox potential and shorter half-life time of •OH than that of SO4•−. (3) The surface charge of Fe3O4@SiO2 is significantly influenced by the acidity and alkalinity of solution. The pKa1 (<0) and pKa2 (9.4) of PMS indicate that PMS primarily exists as HSO5− at a pH below 9.4 and SO52− at a pH above 9.4 [38]. Based on the discussion above, the poor catalytic performance may stem from the electrostatic repulsion between the catalyst and PMS under alkaline conditions.
SO4•− + OH− → SO42− + •OH
SO4•− + H2O → SO42− + •OH + H+
The effect of the reaction temperature on the degradation efficiency of CIP in Fe3O4@SiO2/PMS system was studied. As observed from Figure 11a,b, the CIP removal efficiency increased from 51.8% to 69.2%, 82.8%, and 89.4%, along with the reaction rate constant rising from 0.006 to 0.018, 0.028, and 0.052 min−1, as the temperature climbed from 5 °C to 15 °C, 25 °C, and 35 °C, respectively. The results demonstrated that raising reaction temperatures was beneficial to CIP removal because the higher temperature could accelerate the decomposition of PMS molecules to produce more free radicals. Moreover, the increasing temperature will trigger more effective collision between molecules in the solution, thus increasing the reaction rate in the system [39]. Based on the Arrhenius equation (Equation (6) [40], the activation energy (Ea) of CIP degradation in the Fe3O4@SiO2/PMS system could be calculated as 49.39 kJ/mol (Figure 11c), which was higher than most diffusion-controlled reactions (10–13 kJ/mol) [41]). The findings revealed the intrinsic reaction rate on the catalyst surface rather than the rate of mass transfer, which determine the apparent reaction rate for CIP degradation [42].
Ln k obs = ln A - E a R ( 1 T )

3.5. ROS in Fe3O4@SiO2-PMS System

SO4•− and •OH are the two most prevalent free radicals generated in the catalytic oxidation process of transition metals. Due to the different reaction rates with SO4•− and •OH [43], tert-butanol (TBA) and ethanol (EtOH) have been broadly utilized to identify the role of SO4•− and •OH on CIP removal. EtOH could quench both SO4•− and •OH with the reaction rate constants being (1.6–7.7) × 107 M−1s−1 and (1.2–2.8) × 109 M−1s−1, respectively, while TBA frequently functions as an effective •OH scavenger owing to its faster reaction with •OH (k•OH = (3.8–7.6) × 108 M−1s−1) than SO4•− (kSO4•− = (4.0–9.1) × 105 M−1s−1) [44]. As shown in Figure 12a, the CIP removal efficiency decreased from 82.8% to 61.2% when 50 mM TBA was added, indicating the quenching of •OH generated in the system. The removal efficiency of CIP sharply dropped from 61.2% to 23.8% in the presence of 50 mM EtOH, revealing that SO4•− and •OH produced in the Fe3O4@SiO2/PMS system could be effectively quenched by 50 mM EtOH. The results demonstrated that SO4•− and •OH were simultaneously produced in the system, where SO4•− played the dominant role in CIP degradation over •OH during the oxidation process.
The electron paramagnetic resonance (EPR) experiments with DMPO as the spin-trapping agents were conducted to further examine the free radicals formed in the system [45]. As depicted in Figure 12b, there was no obvious signal when using PMS alone, indicating that PMS could not be decomposed to produce free radicals in the absence of the catalyst. A DMPO-OH signal with the peak intensity ratio of 1:2:2:1 was detected (aN = aH = 14.9 G) in the Fe3O4@SiO2/PMS system, verifying the formation of •OH [46]. The signal of DMPO-SO4 was also discerned at the same time (aN = 13.2 G, aH = 9.6 G, aH = 1.48 G and aH = 0.78 G), further supporting the production of SO4•− in the system [47].

3.6. Possible Activation Mechanisms

Based on the aforementioned discussions, a possible catalytic mechanism of PMS activation by Fe3O4@SiO2 was depicted in Figure 13. The Fe(II) species in the catalyst could directly activate PMS to produce SO4•− and •OH (Equations (7) and (8)), which synergistically contributed to the oxidative degradation of CIP. Then, Fe(III) generated in the reaction system needed to be converted into Fe(II) through reaction with PMS, accompanied by the formation of SO5•− and H+ (Equation (8)). Subsequently, Fe(II) was again involved in the activation of PMS.
Fe(II) + HSO5− → Fe(III) + SO4•− + OH−
Fe(II) + HSO5− → Fe(III) + SO42− + •OH
Fe(III) + HSO5− → Fe(II) + SO5•− + H+

4. Conclusions

In this study, a magnetic heterogeneous catalyst (Fe3O4@SiO2) was successfully prepared via hydrothermal reduction treatment by using drinking water iron-rich sludge and utilized as an PMS activator for CIP removal. Characterizations analysis confirmed that the iron-rich sludge after hydrothermal reduction treatment could be converted to Fe3O4@SiO2 composite with higher specific surface area and magnetic property. The Fe3O4@SiO2 catalyst presented excellent catalytic performance on PMS activation for the CIP degradation, and 82.8% of CIP (10 mg/L) could be removed within 60 min with a dosage of 200 mg/L Fe3O4@SiO2 and 0.5 g/L PMS. In addition, acidic conditions and high temperatures were beneficial to the catalytic reaction in the Fe3O4@SiO2/PMS system. Both quenching tests and EPR analysis illustrated that SO4•− and •OH generated in the system simultaneously participated in the degradation of CIP, and SO4•− played the dominant role during the reaction. This study provides a promising and eco-friendly approach to converting waste iron-rich sludge into an efficient heterogeneous catalyst, which has potential applications in decontamination of the refractory organic pollutants through PMS-AOPs.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/su142416419/s1, Figure S1: XPS survey and spectra of Fe 2p peaks for iron-rich sludge; Figure S2: The EDS mapping images of C, O, Fe, Si, and Ca; Table S1: The calculation of adsorption capacity; Table S2: Elemental Composition of iron sludge by EDS; Text S1: The calculation of adsorption capacity.

Author Contributions

Conceptualization, S.Z.; methodology, S.Z.; software, S.Z.; formal analysis, Z.W. and X.L.; investigation, Z.W.; resources, H.L. (Haojin Luo); data curation, H.L (Haojie Li) and L.W.; writing—original draft preparation, Z.W.; writing—review and editing, S.Z. and J.D.; supervision, S.Z.; project administration, S.Z. and Z.T.; funding acquisition, S.Z. and J.D. All authors have read and agreed to the published version of the manuscript.

Funding

This research is supported by National Natural Science Foundation of China (51978618, 51508509), Natural Science Foundation of Zhejiang Province (LY21E080018, LY18E080036), Foundation of Key Laboratory of Yangtze River Water Environment and Ministry of Education (Tongji University), China (YRWEF201901) and State Key Laboratory of Pollution Control and Resource Reuse Foundation (No. PCRRF21027).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Conflicts of Interest

The authors declare no conflict of interest.

References

  1. Li, L.; He, J.; Xin, X.; Wang, M.; Xu, J.; Zhang, J. Enhanced bioproduction of short-chain fatty acids from waste activated sludge by potassium ferrate pretreatment. Chem. Eng. J. 2018, 332, 456–463. [Google Scholar] [CrossRef]
  2. Zhao, J.; Li, A.; Wang, H. Study on the feasibility and stability of drinking water treatment sludge (DWTS)@zeolite to remove phosphorus from constructed wetlands. J. Environ. Chem. Eng. 2022, 10, 108713. [Google Scholar] [CrossRef]
  3. Huang, P.; Liang, Z.; Zhao, Z.; Cui, F. Synthesis of hydrotalcite-like compounds with drinking water treatment residuals for phosphorus recovery from wastewater. J. Clean. Prod. 2021, 301, 126976. [Google Scholar] [CrossRef]
  4. Fytili, D.; Zabaniotou, A. Utilization of sewage sludge in EU application of old and new methods—A review. Renew. Sust. Energy Rev. 2008, 12, 116–140. [Google Scholar] [CrossRef]
  5. Li, Z.; Gong, M.; Wang, M.; Feng, A.; Wang, L.; Ma, P.; Yuan, S. Influence of AlCl3 and oxidant catalysts on hydrogen production from the supercritical water gasification of dewatered sewage sludge and model compounds. Int. J. Hydrog. Energy 2021, 46, 31262–31274. [Google Scholar] [CrossRef]
  6. Li, D.; Zhuge, Y.; Liu, Y.; Pham, P.N.; Zhang, C.; Duan, W.; Ma, X. Reuse of drinking water treatment sludge in mortar as substitutions of both fly ash and sand based on two treatment methods. Constr. Build. Mater. 2021, 277, 122330. [Google Scholar] [CrossRef]
  7. Jung, K.W.; Hwang, M.J.; Park, D.S.; Ahn, K.H. Comprehensive reuse of drinking water treatment residuals in coagulation and adsorption processes. J. Environ. Manag. 2016, 181, 425–434. [Google Scholar] [CrossRef]
  8. Castaldi, P.; Silvetti, M.; Garau, G.; Demurtas, D.; Deiana, S. Copper(II) and lead(II) removal from aqueous solution by water treatment residues. J. Hazard. Mater. 2015, 283, 140–147. [Google Scholar] [CrossRef]
  9. Gao, J.; Zhao, J.; Zhang, J.; Li, Q.; Gao, J.; Cai, M.; Zhang, J. Preparation of a new low-cost substrate prepared from drinking water treatment sludge (DWTS)/bentonite/zeolite/fly ash for rapid phosphorus removal in constructed wetlands. J. Clean. Prod. 2020, 261, 121110. [Google Scholar] [CrossRef]
  10. Keeley, J.; Smith, A.D.; Judd, S.J.; Jarvis, P. Acidified and ultrafiltered recovered coagulants from water treatment works sludge for removal of phosphorus from wastewater. Water Res. 2016, 88, 380–388. [Google Scholar] [CrossRef] [Green Version]
  11. Kuster, A.C.; Huser, B.J.; Thongdamrongtham, S.; Padungthon, S.; Junggoth, R.; Kuster, A.T. Drinking water treatment residual as a ballast to sink Microcystis cyanobacteria and inactivate phosphorus in tropical lake water. Water Res. 2021, 207, 117792. [Google Scholar] [CrossRef] [PubMed]
  12. Zhang, H.; Liu, X.; Lin, C.; Li, X.; Zhou, Z.; Fan, G.; Ma, J. Peroxymonosulfate activation by hydroxylamine-drinking water treatment residuals for the degradation of atrazine. Chemosphere 2019, 224, 689–697. [Google Scholar] [CrossRef] [PubMed]
  13. Sun, J.; Shen, C.H.; Guo, J.; Guo, H.; Yin, Y.F.; Xu, X.J.; Fei, Z.H.; Liu, Z.T.; Wen, X.J. Highly efficient activation of peroxymonosulfate by Co3O4/Bi2WO6 p-n heterojunction composites for the degradation of ciprofloxacin under visible light irradiation. J. Colloid. Interface Sci. 2021, 588, 19–30. [Google Scholar] [CrossRef] [PubMed]
  14. Wang, Q.; Lu, J.; Jiang, Y.; Yang, S.; Yang, Y.; Wang, Z. FeCo bimetallic metal organic framework nanosheets as peroxymonosulfate activator for selective oxidation of organic pollutants. Chem. Eng. J. 2022, 443, 1385–8947. [Google Scholar] [CrossRef]
  15. Deng, J.; Ye, C.; Cai, A.; Huai, L.; Zhou, S.; Dong, F.; Li, X.; Ma, X. S-doping α-Fe2O3 induced efficient electron-hole separation for enhanced persulfate activation toward carbamazepine oxidation: Experimental and DFT study. Chem. Eng. J. 2021, 420, 129863. [Google Scholar] [CrossRef]
  16. Zou, Z.; Huang, X.; Guo, X.; Jia, C.; Li, B.; Zhao, E.; Wu, J. Efficient degradation of imidacloprid in soil by thermally activated persulfate process: Performance, kinetics, and mechanisms. Ecotoxicol. Environ. Saf. 2022, 241, 113815. [Google Scholar] [CrossRef]
  17. Xu, L.; Wang, X.; Sun, Y.; Gong, H.; Guo, M.; Zhang, X.; Meng, L.; Gan, L. Mechanistic study on the combination of ultrasound and peroxymonosulfate for the decomposition of endocrine disrupting compounds. Ultrason. Sonochem. 2020, 60, 104749. [Google Scholar] [CrossRef] [PubMed]
  18. Guerra-Rodríguez, S.; Rodríguez, E.; Moreno-Andrés, J.; Rodríguez-Chueca, J. Effect of the water matrix and reactor configuration on Enterococcus sp. inactivation by UV-A activated PMS or H2O2. J. Water Process Eng. 2022, 47, 102740. [Google Scholar] [CrossRef]
  19. Zeng, H.; Zhu, H.; Deng, J.; Shi, Z.; Zhang, H.; Li, X.; Deng, L. New insight into peroxymonosulfate activation by CoAl-LDH derived CoOOH: Oxygen vacancies rather than Co species redox pairs induced process. Chem. Eng. J. 2022, 442, 136251. [Google Scholar] [CrossRef]
  20. Qin, Y.; Li, G.; Gao, Y.; Zhang, L.; Ok, Y.S.; An, T. Persistent free radicals in carbon-based materials on transformation of refractory organic contaminants (ROCs) in water: A critical review. Water Res. 2018, 137, 130–143. [Google Scholar] [CrossRef]
  21. Wang, Q.; Jiang, Y.; Yang, S.; Lin, J.; Lu, J.; Song, W.; Zhu, S.; Wang, Z. Selective degradation of parachlorophenol using Fe/Fe3O4@CPPy nanocomposites via the dual nonradical/radical peroxymonosulfate activation mechanisms. Chem. Eng. J. 2022, 445, 136806. [Google Scholar] [CrossRef]
  22. Liang, J.; Xu, X.; Zhong, Q.; Xu, Z.; Zhao, L.; Qiu, H.; Cao, X. Roles of the mineral constituents in sludge derived biochar in persulfate activation for phenol degradation. J. Hazard. Mater. 2020, 398, 122861. [Google Scholar] [CrossRef] [PubMed]
  23. Dong, X.; Ren, B.; Sun, Z.; Li, C.; Zhang, X.; Kong, M.; Zheng, S.; Dionysiou, D.D. Monodispersed CuFe2O4 nanoparticles anchored on natural kaolinite as highly efficient peroxymonosulfate catalyst for bisphenol A degradation. Appl. Catal. B 2019, 253, 206–217. [Google Scholar] [CrossRef]
  24. Lin, R.; Stuckman, M.; Howard, B.H.; Bank, T.L.; Roth, E.A.; Macala, M.K.; Lopano, C.; Soong, Y.; Granite, E.J. Application of sequential extraction and hydrothermal treatment for characterization and enrichment of rare earth elements from coal fly ash. Fuel 2018, 232, 124–133. [Google Scholar] [CrossRef]
  25. Lu, P.; Lin, K.; Gan, J. Enhanced ozonation of ciprofloxacin in the presence of bromide: Kinetics, products, pathways, and toxicity. Water Res. 2020, 183, 116105. [Google Scholar] [CrossRef]
  26. Nadar, A.; Banerjee, A.M.; Pai, M.R.; Meena, S.S.; Pai, R.V.; Tewari, R.; Yusuf, S.M.; Tripathi, A.K.; Bharadwaj, S.R. Nanostructured Fe2O3 dispersed on SiO2 as catalyst for high temperature sulfuric acid decomposition—Structural and morphological modifications on catalytic use and relevance of Fe2O3-SiO2 interactions. Appl. Catal. B 2017, 217, 154–168. [Google Scholar] [CrossRef]
  27. Fang, G.; Gao, J.; Dionysiou, D.D.; Liu, C.; Zhou, D. Activation of persulfate by quinones: Free radical reactions and implication for the degradation of PCBs. Environ. Sci. Technol. 2013, 47, 4605–4611. [Google Scholar] [CrossRef]
  28. Deng, J.; Cheng, Y.; Lu, Y.; Crittenden, J.C.; Zhou, S.; Gao, N.; Li, J. Mesoporous manganese Cobaltite nanocages as effective and reusable heterogeneous peroxymonosulfate activators for Carbamazepine degradation. Chem. Eng. J. 2017, 330, 505–517. [Google Scholar] [CrossRef]
  29. Li, X.; Liu, X.; Lin, C.; Zhang, H.; Zhou, Z.; Fan, G.; He, M.; Ouyang, W. Activation of peroxymonosulfate by magnetic catalysts derived from drinking water treatment residuals for the degradation of atrazine. J. Hazard. Mater. 2019, 366, 402–412. [Google Scholar] [CrossRef]
  30. Zhu, S.; Wang, Z.; Ye, C.; Deng, J.; Ma, X.; Xu, Y.; Wang, L.; Tang, Z.; Luo, H.; Li, X. Magnetic Co/Fe nanocomposites derived from ferric sludge as an efficient peroxymonosulfate catalyst for ciprofloxacin degradation. Chem. Eng. J. 2022, 432, 134180. [Google Scholar] [CrossRef]
  31. Luo, J.; Bo, S.; Qin, Y.; An, Q.; Xiao, Z.; Zhai, S. Transforming goat manure into surface-loaded cobalt/biochar as PMS activator for highly efficient ciprofloxacin degradation. Chem. Eng. J. 2020, 395, 125063. [Google Scholar] [CrossRef]
  32. Ding, Y.; Wang, X.; Fu, L.; Peng, X.; Pan, C.; Mao, Q.; Wang, C.; Yan, J. Nonradicals induced degradation of organic pollutants by peroxydisulfate (PDS) and peroxymonosulfate (PMS): Recent advances and perspective. Sci. Total Environ. 2021, 765, 142794. [Google Scholar] [CrossRef] [PubMed]
  33. Olmez-Hanci, T.; Arslan-Alaton, I. Comparison of sulfate and hydroxyl radical based advanced oxidation of phenol. Chem. Eng. J. 2013, 224, 10–16. [Google Scholar] [CrossRef]
  34. Chen, M.M.; Niu, H.Y.; Niu, C.G.; Guo, H.; Liang, S.; Yang, Y.Y. Metal-organic framework-derived CuCo/carbon as an efficient magnetic heterogeneous catalyst for persulfate activation and ciprofloxacin degradation. J. Hazard. Mater. 2022, 424, 127196. [Google Scholar] [CrossRef] [PubMed]
  35. Zheng, X.; Niu, X.; Zhang, D.; Lv, M.; Ye, X.; Ma, J.; Lin, Z.; Fu, M. Metal-based catalysts for persulfate and peroxymonosulfate activation in heterogeneous ways: A review. Chem. Eng. J. 2022, 429, 132323. [Google Scholar] [CrossRef]
  36. Lai, B.; Chen, Z.; Zhou, Y.; Yang, P.; Wang, J.; Chen, Z. Removal of high concentration p-nitrophenol in aqueous solution by zero valent iron with ultrasonic irradiation (US-ZVI). J. Hazard. Mater. 2013, 250–251, 220–228. [Google Scholar] [CrossRef] [PubMed]
  37. Fadaei, S.; Noorisepehr, M.; Pourzamani, H.; Salari, M.; Moradnia, M.; Darvishmotevalli, M.; Mengelizadeh, N. Heterogeneous activation of peroxymonosulfate with Fe3O4 magnetic nanoparticles for degradation of Reactive Black 5: Batch and column study. J. Environ. Chem. Eng. 2021, 9, 105414. [Google Scholar] [CrossRef]
  38. Rastogi, A.; Al-Abed, S.R.; Dionysiou, D.D. Sulfate radical-based ferrous–peroxymonosulfate oxidative system for PCBs degradation in aqueous and sediment systems. Appl. Catal. B 2009, 85, 171–179. [Google Scholar] [CrossRef]
  39. Yao, Y.; Lu, F.; Zhu, Y.; Wei, F.; Liu, X.; Lian, C.; Wang, S. Magnetic core-shell CuFe2O4@C3N4 hybrids for visible light photocatalysis of Orange II. J. Hazard. Mater. 2015, 297, 224–233. [Google Scholar] [CrossRef]
  40. Du, R.-L.; Wu, K.; Xu, D.-A.; Chao, C.-Y.; Zhang, L.; Du, X.-D. A modified Arrhenius equation to predict the reaction rate constant of Anyuan pulverized-coal pyrolysis at different heating rates. Fuel Process. Technol. 2016, 148, 295–301. [Google Scholar] [CrossRef]
  41. Chen, Y.; Liu, P.; Zhang, R.; Hu, Y.; Yu, Z. Chemical kinetic analysis of the activation energy of diffusion coefficient of sulfate ion in concrete. Chem. Phys. Lett. 2020, 753, 137596. [Google Scholar] [CrossRef]
  42. Tan, Y.; Li, C.; Sun, Z.; Bian, R.; Dong, X.; Zhang, X.; Zheng, S. Natural diatomite mediated spherically monodispersed CoFe2O4 nanoparticles for efficient catalytic oxidation of bisphenol A through activating peroxymonosulfate. Chem. Eng. J. 2020, 388, 124386. [Google Scholar] [CrossRef]
  43. Chen, G.; Wang, H.; Dong, W.; Ding, W.; Wang, F.; Zhao, Z.; Huang, Y. The overlooked role of Co(OH)2 in Co3O4 activated PMS system: Suppression of Co2+ leaching and enhanced degradation performance of antibiotics with rGO. Sep. Purif. Technol. 2023, 304, 122203. [Google Scholar] [CrossRef]
  44. Wang, A.; Ni, J.; Wang, W.; Liu, D.; Zhu, Q.; Xue, B.; Chang, C.-C.; Ma, J.; Zhao, Y. MOF Derived Co-Fe nitrogen doped graphite carbon@crosslinked magnetic chitosan Micro-nanoreactor for environmental applications: Synergy enhancement effect of adsorption-PMS activation. Appl. Catal. B 2022, 319, 121926. [Google Scholar] [CrossRef]
  45. Li, Y.; Zhang, S.; Qin, Y.; Yao, C.; An, Q.; Xiao, Z.; Zhai, S. Preparation of cobalt/hydrochar using the intrinsic features of rice hulls for dynamic carbamazepine degradation via efficient PMS activation. J. Environ. Chem. Eng. 2022, 10, 108659. [Google Scholar] [CrossRef]
  46. Xie, J.; Lin, R.; Liang, Z.; Zhao, Z.; Yang, C.; Cui, F. Effect of cations on the enhanced adsorption of cationic dye in Fe3O4-loaded biochar and mechanism. J. Environ. Chem. Eng. 2021, 9, 105744. [Google Scholar] [CrossRef]
  47. Xu, M.; Yang, J.; Wang, Y.; Lu, B.; Chen, R.; Liu, H. Novel urchin-like Co5Mn-LDH hierarchical nanoarrays: Formation mechanism and its performance in PMS activation and norfloxacin degradation. Sep. Purif. Technol. 2022, 300, 121822. [Google Scholar] [CrossRef]
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Figure 1. (a) XRD patterns, (b) FTIR spectra, (c) N2 adsorption/desorption isotherms, and (d) BET-specific surface area of Fe3O4@SiO2 catalyst and iron-rich sludge.
Figure 1. (a) XRD patterns, (b) FTIR spectra, (c) N2 adsorption/desorption isotherms, and (d) BET-specific surface area of Fe3O4@SiO2 catalyst and iron-rich sludge.
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Figure 2. (a) SEM and (b) TEM images of iron-rich sludge; (c) SEM and (d) TEM images of Fe3O4@SiO2.
Figure 2. (a) SEM and (b) TEM images of iron-rich sludge; (c) SEM and (d) TEM images of Fe3O4@SiO2.
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Figure 3. (a)TGA−DSC curves of iron-rich sludge in air atmosphere; (b) the magnetic hysteresis loops of Fe3O4@SiO2 catalyst and iron-rich sludge.
Figure 3. (a)TGA−DSC curves of iron-rich sludge in air atmosphere; (b) the magnetic hysteresis loops of Fe3O4@SiO2 catalyst and iron-rich sludge.
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Figure 4. (a) The removal efficiency and (b) adsorption capacity of CIP by iron-rich sludge at different dosages. (Reaction conditions: [CIP]0 = 10 mg/L, T = 25 °C, pH = 7.0 ± 0.2.).
Figure 4. (a) The removal efficiency and (b) adsorption capacity of CIP by iron-rich sludge at different dosages. (Reaction conditions: [CIP]0 = 10 mg/L, T = 25 °C, pH = 7.0 ± 0.2.).
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Figure 5. Surface interaction between the iron-rich sludge and CIP molecule.
Figure 5. Surface interaction between the iron-rich sludge and CIP molecule.
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Figure 6. Comparisons of the CIP degradation with two types of persulfates (PMS and PDS) activated by iron-rich sludge. (Reaction conditions: [CIP]0 = 10 mg/L, [Catalyst] = 200 mg/L, [Persulfate] = 1.6 mM, T = 25 °C, pH = 7.0 ± 0.2.).
Figure 6. Comparisons of the CIP degradation with two types of persulfates (PMS and PDS) activated by iron-rich sludge. (Reaction conditions: [CIP]0 = 10 mg/L, [Catalyst] = 200 mg/L, [Persulfate] = 1.6 mM, T = 25 °C, pH = 7.0 ± 0.2.).
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Figure 7. Effects of (a) iron-rich sludge dosage and (b) PMS concentration on the CIP degradation in iron-rich sludge/PMS system. (Reaction conditions: [CIP]0 = 10 mg/L, [PMS]0 = 0.5 g/L (a), [Catalyst] = 200 mg/L (b), T = 25 °C, pH = 7.0 ± 0.2).
Figure 7. Effects of (a) iron-rich sludge dosage and (b) PMS concentration on the CIP degradation in iron-rich sludge/PMS system. (Reaction conditions: [CIP]0 = 10 mg/L, [PMS]0 = 0.5 g/L (a), [Catalyst] = 200 mg/L (b), T = 25 °C, pH = 7.0 ± 0.2).
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Figure 8. (a) The CIP degradation efficiencies in different systems and (b) comparisons of catalytic activities of two persulfates (PMS and PDS) by Fe3O4@SiO2. (Reaction conditions: [CIP]0 = 10 mg/L, [Catalyst] = 200 mg/L, [PMS, PDS]0 = 1.65 mM, T = 25 °C, pH = 7.0 ± 0.2.).
Figure 8. (a) The CIP degradation efficiencies in different systems and (b) comparisons of catalytic activities of two persulfates (PMS and PDS) by Fe3O4@SiO2. (Reaction conditions: [CIP]0 = 10 mg/L, [Catalyst] = 200 mg/L, [PMS, PDS]0 = 1.65 mM, T = 25 °C, pH = 7.0 ± 0.2.).
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Figure 9. Effects of (a) Fe3O4@SiO2 dosage and (b) PMS concentration on the CIP degradation. (Reaction conditions: [CIP]0 = 10 mg/L, [Catalyst] = 200 mg/L (b), [PMS]0 = 0.5 g/L (a), T = 25 °C, pH = 7.0 ± 0.2.).
Figure 9. Effects of (a) Fe3O4@SiO2 dosage and (b) PMS concentration on the CIP degradation. (Reaction conditions: [CIP]0 = 10 mg/L, [Catalyst] = 200 mg/L (b), [PMS]0 = 0.5 g/L (a), T = 25 °C, pH = 7.0 ± 0.2.).
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Figure 10. Effects of initial solution pH on the CIP degradation by Fe3O4@SiO2/PMS. (Reaction conditions: [CIP]0 = 10 mg/L, [Fe3O4@SiO2] = 200 mg/L, [PMS]0 = 0.5 g/L, T = 25 °C).
Figure 10. Effects of initial solution pH on the CIP degradation by Fe3O4@SiO2/PMS. (Reaction conditions: [CIP]0 = 10 mg/L, [Fe3O4@SiO2] = 200 mg/L, [PMS]0 = 0.5 g/L, T = 25 °C).
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Figure 11. (a) Effects of reaction temperature on the CIP degradation, (b) linear fit for Arrhenius equation, and (c) activation energy by the Fe3O4@SiO2/PMS system. (Reaction conditions: [CIP]0 = 10 mg/L, [Fe3O4@SiO2] = 200 mg/L, [PMS]0 = 0.5 g/L, pH = 7.0 ± 0.2).
Figure 11. (a) Effects of reaction temperature on the CIP degradation, (b) linear fit for Arrhenius equation, and (c) activation energy by the Fe3O4@SiO2/PMS system. (Reaction conditions: [CIP]0 = 10 mg/L, [Fe3O4@SiO2] = 200 mg/L, [PMS]0 = 0.5 g/L, pH = 7.0 ± 0.2).
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Figure 12. (a) Effect of TBA and EtOH on the CIP degradation; (b) EPR spectra of DMPO trapped free radical adducts in Fe3O4@SiO2/PMS system. (Reaction conditions: [CIP]0 = 10 mg/L, [Fe3O4@SiO2] = 200 mg/L, [PMS]0 = 0.5 g/L, [TBA, EtOH] = 50 mM, [DMPO]0 = 20 mM, T = 25 °C, pH = 7.0 ± 0.2).
Figure 12. (a) Effect of TBA and EtOH on the CIP degradation; (b) EPR spectra of DMPO trapped free radical adducts in Fe3O4@SiO2/PMS system. (Reaction conditions: [CIP]0 = 10 mg/L, [Fe3O4@SiO2] = 200 mg/L, [PMS]0 = 0.5 g/L, [TBA, EtOH] = 50 mM, [DMPO]0 = 20 mM, T = 25 °C, pH = 7.0 ± 0.2).
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Figure 13. Catalytic mechanisms in Fe3O4@SiO2/PMS system for CIP degradation.
Figure 13. Catalytic mechanisms in Fe3O4@SiO2/PMS system for CIP degradation.
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Table
Table 1. Physical characteristics of the Fe3O4@SiO2 and iron-rich sludge.
Table 1. Physical characteristics of the Fe3O4@SiO2 and iron-rich sludge.
CatalystsSpecific Surface Area (m2/g)Pore Volume (cm3/g)Pore Size (Å)
Iron-rich sludge21.4860.059133.655
Fe3O4@SiO250.3520.10357.563
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Wang, Z.; Zhu, S.; Deng, J.; Li, H.; Wang, L.; Luo, H.; Tang, Z.; Li, X. Facile Preparation of Fe3O4@SiO2 Derived from Iron-Rich Sludge as Magnetic Catalyst for the Degradation of Organic Contaminants by Peroxymonosulfate Activation. Sustainability 2022, 14, 16419. https://doi.org/10.3390/su142416419

AMA Style

Wang Z, Zhu S, Deng J, Li H, Wang L, Luo H, Tang Z, Li X. Facile Preparation of Fe3O4@SiO2 Derived from Iron-Rich Sludge as Magnetic Catalyst for the Degradation of Organic Contaminants by Peroxymonosulfate Activation. Sustainability. 2022; 14(24):16419. https://doi.org/10.3390/su142416419

Chicago/Turabian Style

Wang, Zhiwei, Shijun Zhu, Jing Deng, Haojie Li, Liang Wang, Haojin Luo, Zehe Tang, and Xueyan Li. 2022. "Facile Preparation of Fe3O4@SiO2 Derived from Iron-Rich Sludge as Magnetic Catalyst for the Degradation of Organic Contaminants by Peroxymonosulfate Activation" Sustainability 14, no. 24: 16419. https://doi.org/10.3390/su142416419

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