Ce-Doped zero-valent iron nanoparticles as a Fenton-like catalyst for degradation of sulfamethazine

Abstract

Ce-Doped zero-valent iron (Ce/Fe) nanoparticles were prepared, characterized and used as a catalyst for degradation of sulfamethazine (SMT) antibiotics in a Fenton-like process. High resolution transmission electron microscopy (TEM), emission scanning electron microscopy (SEM), X-ray diffraction (XRD), X-ray photoelectron spectroscopy (XPS), Raman and Fourier transform infrared (FTIR) spectroscopy were used to characterize the catalyst before and after use. The influencing factors on the degradation of sulfamethazine were determined, including pH value, H₂O₂ dosage, catalyst dosage and temperature. The results showed that Ce/Fe composites exhibited high catalytic activity in a wide pH range (4–7). The removal efficiency of SMT was almost 100% under optimal condition (pH = 6, T = 30 °C, 0.5 g L⁻¹ catalyst and 12 mM H₂O₂). The degradation process of SMT conformed to a first-order kinetic model. The intermediates of SMT degradation were identified by Ion Chromatography (IC), the possible degradation pathway of SMT and the possible catalytic mechanisms of Ce/Fe composites were tentatively proposed.

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1. Introduction

Sulfamethazine (SMT) has been widely used in animal husbandry for its broad antifungal spectrum, low-cost, chemical stability and convenience. Production of sulfamethazine in China has been increasing in recent years. The presence of sulfonamide antibiotics has been frequently detected in aquatic environments.¹

Different technologies have been studied and used to remove antibiotic from water and wastewater. For example, membrane filtration,² activated carbon adsorption,³ advanced oxidation processes (AOPs).^4–8 Heterogeneous Fenton-like process has been receiving increasing attention in recent years to overcome the disadvantages in homogeneous Fenton process, such as: (1) homogeneous Fenton process is highly dependent on pH value, usually it requires pH < 3, which is very narrow for reaction; (2) Fe(II) is easy to precipitate into slurries, which are difficult to remove; (3) iron ions in the solution are difficult to recover. Zero valent iron (nZVI) has been widely studied in degradation of pollutants which has a certain surface area and can provide Fe(II) needed in reaction.^9,10

The mechanism of heterogeneous Fenton-like process is similar to that of the traditional homogeneous Fenton process, which is based on HO˙ and Fe(II). The important step was Fenton process (reaction (1)), that is to say, Fe(II) reacts with H₂O₂ to produce HO˙ through reactions (1)–(4). Fe(III) can be then transferred into Fe(II) with H₂O₂ (reactions (5) and (6)).

Fe²⁺ + H₂O₂ → Fe³⁺ + HO⁻ + HO˙ (1)

Fe²⁺ + HO˙ → Fe³⁺ + HO⁻ (2)

HO˙ + H₂O₂ → H₂O + HO₂˙ (3)

HO₂˙ + H₂O₂ → H₂O + O₂ + HO˙ (4)

Fe³⁺ + H₂O₂ → Fe²⁺ + H⁺ + HO₂˙ (5)

Fe³⁺ + HO₂˙/O₂˙⁻ → Fe²⁺ + H⁺ + O₂ (6)

However, the cycle of Fe(III) and Fe(II) is limited by reactions (5) and (6) for reaction rate constant k₄ = 2 × 10⁻³ M⁻¹ s⁻¹ (reaction (5)) and the concentration of HO₂˙ is low. In addition, Fe⁰ is easy to aggregate during the preparation process,¹¹ which can lead to the decrease of catalytic activity. Many methods have been used to enhance the catalyst activity, such as loading with porous materials to prevent the aggregation of nZVI,^12–14 combining with other oxidation processes (photo, electro, UV, etc.) and adding transition metal¹⁵ which can produce synergistic effect.

Cerium is the most abundant rare earth element, with a high redox potential of E⁰(Ce⁴⁺/Ce³⁺) = 1.84 V, which has been extensively used in treatment of pollutants.^16–18 Ce(III) can facilitate the production of HO₂˙/O₂˙⁻,^19,20 which can enhance the reduction from Fe(III) to Fe(II). The cerium has function of storage of oxygen and can produce O₂˙⁻ which can facilitate the production of HO˙. Xu and Wang²¹ studied the degradation of chlorophenols with Fe⁰/CeO₂ under air-saturated conditions with 48% of mineralization. Ling et al.²² prepared the Ce–Fe bimetallic oxides on graphene and used to adsorb Congo red.

The objective of this study was to prepare Ce/Fe composite catalyst through doping Ce in zero-valent iron nanoparticles, to investigate the performance and mechanism of sulfamethazine (SMT) antibiotics degradation by Fenton-like process using Ce/Fe composite as catalyst.

2. Materials and methods

2.1 Chemicals

Iron(II) sulfate heptahydrate (FeSO₄·7H₂O), cerium(III) nitrate hexahydrate (Ce(NO₃)₃·6H₂O) and potassium borohydride (KBH₄), hydrogen peroxide (30% w/w) and chromatography grade methanol are purchased from Sinopharm Chemical Reagent Co., Ltd. Hydrochloric acid (0.1 M) and sodium hydroxide (0.1 M) were used for pH adjustment. MilliQ water was used throughout the study. Sulfamethazine (C₁₂H₁₄O₂N₄S) (>99%) was purchased from Alfa Aesar. All chemicals were used without further purification.

2.2 Preparation and characterization of Ce/Fe composite

A mixed solution of FeSO₄·7H₂O and Ce(NO₃)₃·6H₂O were added into a four-necked flask container in mole ratio of Fe and Ce = 20 : 1. Temperature was kept at 21 ± 1 °C. After stirring for 15 min, a certain amount of KBH₄ aqueous solution was added continuously into the container at a constant flow rate. The volume ratio of KBH₄ solution to mixed solution was 1 : 1–2.5 : 1, mole ratio was 5 : 1–8 : 1. After 60 min of reaction, nano-sized Ce-doped Fe composites were synthesized. The composites were washed with oxygen-free deionized water for three times, and dried in a vacuum freezing dryer.

X-ray diffraction (XRD) measurement was taken with X-ray power diffractometer (D8-Advance, Bruker, 40 kV and 40 Ma, Cu Kα) at room temperature with 1° min⁻¹ in a range of 10–90°. Emission scanning electron microscope (SEM) was performed using FEI Quanta 200 FEG ESEM instrument of FEI Company with EDX analysis. Samples were prepared with gold and platinum plated on the surface for analysis. High resolution transmission electron microscopies (HRTEM, JEM 2100 and JEOL) operated at 200 kV to observe the microscopic morphology of catalyst loaded in copper grid. Through SEM and TEM, the morphology of catalysts before and after reactions can be observed. The catalysts were prepared in a vacuum freezing dryer.

The Brunauer–Emmett–Teller (BET) specific surface area and Barrett–Joyner–Halenda (BJH) pore size distribution were measured by nitrogen adsorption–desorption isotherm measurements at 77 K on a NOVA 3200e sorptometer and degassing at 428 K.

A physical property measurement system (PPMS, 730T, Lakeshore, USA) was used for magnetization curves measurement.

X-ray photoelectron spectroscopy (XPS) measurements (Thermo Scientific ESCALAB 250Xi) were performed with an Al Kα X-ray (1486.6 eV) source for excitation.

Raman spectrometer (LabRAM HR Evolution of HOEIBA Jobin Yvon Company, French) was used for recording the Raman spectra with a regular model laser operated at wavelength of 532 nm. The laser power was 0.8 mW. The spectra were recorded using a 50× object lens and 600 gr per mm grating.

The infrared spectra were recorded with Nicolet 6700 Fourier transform infrared spectrometer made by Thermo Fisher Scientific. The samples were prepared using powder pressing method in KBr pellet at room temperature.

2.3 Experimental procedures

The experiments were performed in serum bottles (60 mL) that were sealed in a shaking incubator with 160 rpm at a certain temperature in the dark. The reaction suspension containing calculated amount of catalysts and sulfamethazine was prepared, pH was adjusted with HCl (0.1 M) and NaOH (0.1 M). Each experiment was carried out twice, and the results were the average value.

2.4 Analytical methods

The samples were analyzed after filtration with 0.22 μm filter film. The concentration of sulfamethazine was measured using high performance liquid chromatograph (HPLC Agilent 1200) equipped with a diode array detector (DAD) and an XDB-C18 (4.6 × 150 mm) column, the column temperature was 30 °C, the mobile phase was a mixture of distilled water and ethanol (55 : 45 (v/v)) with a flow rate of 1 mL min⁻¹ at the detection wavelength of 255 nm.

Nitrite, nitrate, sulfate, formic acid, oxalic acid, and acetic acid were analyzed by ion chromatography (Dionex model ICS 2100) coupled with a dual-piston pump, a Dionex IonPac AS19/AS11 analytical column (4 mm × 250 mm), an IonPac AG19/AG11 guard column (4 mm × 250 mm), and a DS6 conductivity detector. The eluent solution composed of 3.5 mM Na₂CO₃ and 1.0 mM NaHCO₃ supplying at a rate of 1 mL min⁻¹.

Iron and cerium ions were detected by ICP-MS Thermo ICP-MS iCAPQ made by ThermoFisher under condition of 29 K. The examination standard was JY/T 015-1996.

The value of pH was measured by pH meter (Thermo Orion 8103BN, USA). TOC was determined using Multi 2100TOC/TN analyzer (Analytik Jena AG Corporation).

3. Results and discussion

3.1 The crystal structure of catalyst

The XRD pattern of Ce/Fe composites is shown in Fig. 1. The XRD pattern of Fe⁰ is a hexagonal structure (JCPDS 65-4899) with space group 1m3̄m (2 2 9).²³ The peak at 2θ values of 44.7° corresponded to (1 1 0) crystal planes, respectively. The small peak at 2θ values of 28.6° corresponded to (1 1 1) crystal planes, indicating the existence state of CeO₂ (JCPDS 65-5923). There was also small peak at 35.5° which corresponded to (3 1 1) of Fe₃O₄ (JCPDS 65-3107).

Fig. 1 XRD patterns of Ce/Fe composites before reaction.

3.2 Morphology of catalyst

The SEM and TEM micrographs of the catalysts in different magnification times are presented in Fig. 2 and 3. As shown in Fig. 2, we can see that Ce/Fe composites were spherical with diameter of about 100 nm. It was larger than Fe⁰ composites with diameter of around 20 nm (Fig. 3b). Compared Fig. 3a and b, we can see that Ce doped Fe⁰ can effectively inhibit the aggregation of Fe⁰. The aggregation of Fe⁰ will lead to the decrease of surface area and reactivity.¹¹ The EDX analysis of Ce/Fe composites indicated the element content of Fe and Ce. Fig. 3a-1 shows that the atomic ratio of Fe and Ce was from 59.62% to 1.09%. Fig. 3c shows the morphology of catalysts after reaction, it can be seen that catalysts displayed rod-like shape as reported in previous study,²⁴ suggesting the corrosion of catalyst by H₂O₂ and the formation of iron oxide.

Fig. 2 SEM micrographs (a) Ce/Fe composites, 100k; (b), (c) Ce/Fe composites, 250k; (d) single particle of Ce/Fe, 300k.

Fig. 3 TEM micrographs (a) Ce/Fe composites before reaction; (a-1) EDX data of Ce/Fe composites; (b) Fe⁰ composites; (c) Ce/Fe composites after reaction.

3.3 BET surface area of Ce/Fe composites

The nitrogen adsorption/desorption isotherm and pore size distribution of Ce/Fe composites are showed in Fig. S1.† The specific surface areas (S_BET) and pore volume were 24.04 m² g⁻¹ and 0.098 cm³ g⁻¹.

According to BDDT (Brunauer–Deming–Deming–Teller), the BET isotherms of Ce/Fe composites are type IV with H3 type hysteresis loops, indicating that Ce/Fe composites were mesoporous material (Fig. S1a†). Fig. S1b† shows the pore size distribution and the pore diameter was 2.05 nm, belonging to mesopore. The S_BET and pore volume of Fe⁰ prepared with the same method were 3.89 m² g⁻¹ and 0.01 cm³ g⁻¹.²¹ Therefore, the addition of Ce could increase the S_BET and pore volume, which are favorable to the degradation of organic pollutants.

3.4 XPS analysis

To find out the change of Ce/Fe composites before and after reaction, the survey scan and high resolution scan of Fe 2p region and Ce 3d region was carried out (Fig. 4). The information of surface characteristics of Fe and Ce elements could be obtained through XPS analysis.

Fig. 4 XPS spectra of Ce/Fe composites (a) survey scan; (b) high resolution scan of Fe 2p region; (c) high resolution scan of Ce 3d region.

The calibrated binding energy (BE) for Fe 2p and Ce 3d based on carbon signal at 284.8 eV could be obtained. Fig. 4b shows the spin–orbit doublet of Fe 2p spectra before and after reaction. The binding energy of 724.6 eV, 711.0 eV and 706.6 eV corresponded to Fe 2p_1/2, Fe 2p_3/2 and Fe(0) 2p_3/2, respectively.^17,25 The peaks at 706.6, 710.7, and 712.2 eV can be ascribed to Fe(0), Fe(II) and Fe(III), respectively.²⁶ The peak at 706.6 eV was too small to be detected after the reaction, compared with the peak at 706.6 eV before reaction, revealing that oxidation happened on Ce/Fe composites after reaction. The existence of Fe(III) and Fe(II) before reaction indicated that the surface of Fe(0) was oxidized in the process of storage and core–shell structure was formed. The peak area ratio of Fe(II) to Fe(III) was 1.7 and 1 before and after reaction, respectively. Combining the analysis of XRD, TEM, and XPS, we supposed that the thin layer of iron oxide was combination of γ-Fe₂O₃ and Fe₃O₄ formed outside the Ce/Fe composites.

The high-resolution spectra of spin–orbital doublets (Ce 3d_3/2 and Ce 3d_5/2) in Ce/Fe composites are shown in Fig. 4c. From Fig. 4c we can see the existence of peaks of Ce(IV), Ce(III) and Ce(0) in Ce/Fe composites before reaction.²⁷ While after the reaction, the peak area of Ce(III) increased with decrease of peak area of Ce(IV), which can be ascribed to the electron transfer between Ce(III) and Ce(IV).²¹ The disappearance of Ce(0) may also resulted in the increase of Ce(III) when the electron transfer from Ce to H₂O₂ and O₂ (reactions (7) and (8)). The increase of Ce(III) can increase the chemisorbed oxygen and vacancies on the surface of catalyst.^18,28 The chemisorbed oxygen is the most important oxygen in the process of oxidation, which can increase the activity of catalysts¹⁷ as described in the reactions (9) and (10).

4Ce⁰ + 3O₂ + 6H₂O → 4Ce³⁺ + 12OH⁻ (7)

2Ce⁰ + 3H₂O₂ + 6H⁺ → 2Ce³⁺ + 6H₂O (8)

Ce³⁺ + O₂ → Ce⁴⁺ + ˙O₂⁻ (9)

˙O₂⁻ + H⁺ → 2HO˙ (10)

3.5 FTIR and Raman spectroscopy

FTIR spectra of Ce/Fe composites before and after reaction are shown in Fig. 5. The band centered at 3431 cm⁻¹, 1636 cm⁻¹ and 1400 cm⁻¹ were aroused from the adsorbed water molecule,²⁹ HO˙ group in the deeper structure of iron oxide and H₂O₂,³⁰ respectively. Some extra peaks at 1053 cm⁻¹, 600 cm⁻¹ and 440 cm⁻¹ displayed in lower wave numbers. These bands may originate from chemical bond Fe–O in magnetite (Fe₃O₄), maghemite (γ-Fe₂O₃) or lepidocrocite (γ-FeO(OH)). After reaction, a new small peak at 709 cm⁻¹ can be ascribed to nitrate group³¹ which is the by-product of system. The new peak at 1173 and 1119 cm⁻¹ are corresponding to γ-FeO(OH) and akaganeite (β).²⁹ According to Turković et al.,³² characteristic absorption wavelength of γ_Ce–O is about 475 cm⁻¹.

Fig. 5 FTIR spectra of Ce/Fe composites before and after reaction.

Fig. 6 shows the Raman spectra of Ce/Fe composites before and after reaction. The band centered at 209 cm⁻¹, 274 cm⁻¹, 395 and 575 cm⁻¹ represented Fe–O in composites before the reaction.^33,34 After reaction, all bands were enforced, indicating the corrosion of catalysts in the reaction. However, there was no bands information of Ce element in the Raman spectra centered at 460 cm⁻¹ which represented Ce⁴⁺,³⁵ 2134 cm⁻¹ and 2269 cm⁻¹ which represented vibration Raman peaks of Ce(III),³⁶ which may be caused by low content of Ce.

Fig. 6 Raman spectra of Ce/Fe composites before and after reaction.

3.6 Magnetization characterization of catalyst

Fig. 7(a) and (b) shows the effect of exterior magnetic field on Ce/Fe composites. It can be seen that the composites can be easily homogeneously separated. Vibrating-sample magnetometer was used to study the behaviour of Ce/Fe composites at room temperature in the field range of −10 kOe < H < 10 kOe. From Fig. 7c, the saturated magnetization (M_s) of Ce/Fe composites was 98.65 A m² kg⁻¹, and the coercivity was 110.06 G. Jabeen et al.³⁷ prepared Fe⁰ through co-precipitation method without doped Ce, the M_s value of Fe⁰ was 173 A m² kg⁻¹. The decrease in M_s value of Ce/Fe composites may ascribe to the loading of Ce and the surface oxidation. This magnetic properties mean that Ce/Fe nanocomposites can be easily separated in the practical application.

Fig. 7 Digital photo of Ce/Fe composites suspension without (a) and with (b) exterior magnetic field; (c) magnetic hysteresis curve of the Ce/Fe composites.

3.7 Effect of different catalysts on SMT degradation

The effect of different catalysts, including Ce/Fe composites and Fe⁰ on SMT degradation in Fenton-like process was investigated. The results are shown in Fig. 8.

Fig. 8 Degradation of SMT with different catalysts (SMT = 20 mg L⁻¹, 30 °C, pH = 7, 1 g L⁻¹ catalyst, 8 mM H₂O₂).

Fig. 8 shows the results of SMT degradation using different catalysts under condition of 30 °C, pH = 7, 1 g L⁻¹ catalyst and 8 mM H₂O₂. We conducted control experiments using catalyst of Fe⁰ with and without Ce doped to identify the validation of Ce/Fe composites. The experiments using Ce/Fe composites as catalyst without H₂O₂ and only 8 mM H₂O₂ were also performed. As shown in Fig. 8, we can see that when pH was 7, the removal efficiency of SMT was only 22% for Fe⁰ as catalyst. When Ce was doped, the removal efficiency increased to almost 67% at pH = 7, suggesting that Ce addition could improve the activity of catalyst. Fig. 8 also showed that only 4–5% SMT was removed if only using catalyst or H₂O₂, owing to the adsorption of catalyst and the oxidation by only H₂O₂. The concentration of dissolved Fe and Ce when using Ce/Fe composites as catalyst in the final solution after 24 h was 0.94 mg L⁻¹ and 0.27 mg L⁻¹, respectively. The leaching rate of Fe and Ce was 0.2% and 1.1%. We carried out experiments with 1 mg L⁻¹ Fe(II) and 0.3 mg L⁻¹ Ce(III) as catalyst under condition of 30 °C, pH = 7 and 8 mM H₂O₂, the removal efficiency of SMT was only 10%. The results showed that the Ce/Fe composites are potential catalyst in heterogeneous Fenton-like process.

3.8 Degradation of SMT under different conditions

The influence of different factors on SMT degradation using Ce/Fe composites as catalyst was studied, including pH value, temperature, H₂O₂ concentration and catalysts dosage. The results are shown in Fig. 9.

Fig. 9 Effect of environmental factors on SMT (20 mg L⁻¹) degradation (a) pH value, 1 g L⁻¹ catalyst, 8 mM H₂O₂; (b) H₂O₂ concentration, 1 g L⁻¹ catalyst; (c) catalyst dosage, 8 mM H₂O₂ (d) temperature, 0.5 g L⁻¹ catalyst, 12 mM H₂O₂. Except for the investigated parameter, other parameters fixed on pH = 7, 30 °C.

Fig. 9a shows the influence of pH values, indicating that removal efficiency of SMT increased when pH decreased. However, there was no obvious difference at pH = 5 or 6, the removal efficiency of SMT reached 100% at 5 min when pH = 5, at 10 min when pH = 6. When pH increased to 8, there was almost no degradation of SMT. The reasons may be: (1) when pH was high, Fe(II) and Fe(III) can be easily transformed to Fe(OH)(H₂O)₅^+/2+, leading to less amount of hydroxyl radicals; (2) the decomposition of H₂O₂ itself was accelerated at high pH values; (3) the soluble iron ions can easily precipitate and the surface of Fe⁰ was passivated in alkaline condition.^38,39 This is accordance with the results obtained by other researchers.⁴⁰

Fig. 9b shows the influence of concentration of H₂O₂. It can be seen that the removal efficiency of SMT increased to 74% when H₂O₂ concentration increased to 12 mM. However, when H₂O₂ concentration further increased from 12 mM to 16 mM, the removal efficiency of SMT decreased to 54% gradually, which accorded with the results of Ma et al.⁴¹ They found that hydroxyl radicals can be consumed by excess H₂O₂, leading to the decrease of HO˙ in the solution (reaction (3)). In addition, excess HO˙ can react with Fe(II), leading to the decrease of Fe(II) (reaction (11)).

Fe²⁺ + HO˙ → Fe³⁺ + HO⁻ (11)

Fig. 10c shows the influence of different amount of catalyst. It can be seen that when catalyst dosage increased from 0.1 g L⁻¹ to 0.5 g L⁻¹, the removal efficiency of SMT increased from 67% to 82%. When it further increased to 1 g L⁻¹ and 1.5 g L⁻¹, the removal efficiency of SMT decreased to 67% and 45%, respectively. The reason may be that excess of ferrous irons can consume HO˙ to form ferric ions as reaction (2).

Fig. 10 (a) Evolution of SMT and TOC (b) evolution of inorganic ions during SMT degradation process (SMT = 20 mg L⁻¹, 30 °C, pH = 6, 0.5 g L⁻¹ catalyst, 12 mM H₂O₂).

Fig. 10d shows the influence of temperature. We can see that the removal efficiency of SMT increased as temperature increased, indicating that SMT degradation process is endothermic reaction.

3.9 SMT degradation kinetics

The first-order degradation kinetic was used to simulate the SMT degradation process at different conditions, the results showed a good linear relationship (Fig. S2†). It can be seen from Fig. S2(a)† that when pH = 5, the rate constant (k) was highest (0.719 min⁻¹). When pH increased, the k values decreased. From Fig. S2(b)† we can see that when H₂O₂ concentration increased from 4 mM to 12 mM, the rate constant increased from 0.0765 to 0.243 min⁻¹. Addition of 16 mM H₂O₂ inhibited the process, with k of 0.122 min⁻¹. As shown in Fig. S2(c),† when the catalyst dosage was 0.5 g L⁻¹, k was 0.202 min⁻¹ which was highest among catalyst dosage of 0.1–1.5 g L⁻¹. Fig. S2(d)† shows that k increased with increase of temperature.

3.10 SMT degradation pathway

The removal of SMT and TOC during the degradation process was 100% and 62.2%, as shown in Fig. 10a. In previous study by Liu et al.,⁴² TOC removal efficiency was about 15% in the process of Fe(II) enhanced sulfamethazine degradation. The main inorganic ions (NO₃⁻, NO₂⁻ and SO₄²⁻) produced at different time during Fenton-like degradation of SMT were determined using ion chromatography (Fig. 10b). We can see that their concentration changed with SMT degradation. The concentration of SO₄²⁻ increased rapidly at first 2 min, due to the attack of sulfonamide bond S–N in SMT by ˙OH. Nitrogen atoms were released from the pyrimidine ring breakage and azo bond cleavage, generating NO₃⁻ and NO₂⁻ after first 5 min, indicating that the C–N bond was difficult to be broken than S–N bond.

To analyze the final inorganic and organic products, the mass balance was performed. The addition of SMT (C₁₂H₁₄N₄O₂S) was 0.0719 mM corresponding the original addition of 20 mg L⁻¹. From the Fig. 10b, we can see that S in the form of sulfate is about 0.02 mM which is only 32% of the addition of S. The N in the form of nitrate and nitrite ions is 0.16 mM which is 65% of the total nitrogen. This results indicated that the intermediates containing S and N resulted in the final TOC. In our previous studies,^43–46 the intermediate products were examined suing GC-MS analyses.

Some small molecular organic acids, such as formic acid, acetic acid and oxalic acid were detected. Acetic acid was oxidized to formic and oxalic acids, which are ultimate products that directly evolved to CO₂. Moreover, all these acids might form Fe(III)–carboxylate complexes since iron ions were released to the medium.⁴⁷

3.11 Possible catalytic mechanism of catalyst

The analysis with XRD, XPS, TEM indicated that the existence of CeO₂, Fe₃O₄, Fe⁰ and Ce⁰ on the fresh catalyst surface. The reduction potential of Fe(0) and Fe(II) is −0.447 V. When there are dissolved oxygen and hydrogen peroxide, the electrons can transfer to Fe(0) from O₂ and H₂O₂ to form Fe(II) (reactions (14) and (15)). The reduction potential of Ce(0) and Ce(III) is −2.336 V. In the same way, Ce(III) can be formed through the electron transfer from Ce(0) to O₂ (reactions (12) and (13)).

4Ce⁰ + 3O₂ + 6H₂O → 4Ce³⁺ + 12OH⁻ (12)

2Ce⁰ + 3H₂O₂ + 6H⁺ → 2Ce³⁺ + 6H₂O (13)

2Fe⁰ + O₂ + 2H⁺ → Fe²⁺ + H₂O₂ (14)

Fe⁰ + H₂O₂ + 2H⁺ → Fe²⁺ + H₂O (15)

The addition of cerium can enhance the catalyst activity as follows: firstly, Ce(IV) can be reduced to Ce(III) by H₂O₂ (reactions (16) and (17)) and produce HO₂˙/O₂˙⁻ which can enhance the cycle of Fe(III) and Fe(II);^19,20 secondly, the reduction potential of Ce(IV)/Ce(III) and Fe(III)/Fe(II) are 1.44 V and 0.77 V, respectively, electrons can be transferred from Fe(0) to Ce(IV) and produced Fe(II), which can facilitate the potential of catalytic activity (reactions (18) and (19)); thirdly, the addition of ceria can make oxygen store on the surface of the catalyst,⁴⁸ which can facilitate the production of high active free radical ˙O₂⁻ and HO˙ (reactions (9) and (10)).^28,49

Ce⁴⁺ + H₂O₂ → HO₂˙/O₂˙⁻ + H⁺ + Ce³⁺ (16)

Ce⁴⁺ + HO₂˙/O₂˙⁻ → Ce³⁺ + H⁺ + O₂ (17)

2Ce⁴⁺ + Fe⁰ → 2Ce³⁺ + Fe²⁺ (18)

Ce⁴⁺ + Fe²⁺ → Ce³⁺ + Fe³⁺ (19)

In summary, the cycle of Ce(III) and Ce(IV) can be achieved through reactions ((9), (10) and (16)–(19)), the addition of cerium can facilitate the production of HO₂˙, O₂˙⁻ and HO˙, HO₂˙/O₂˙⁻ can also enhance the reduction of Fe(III) to Fe(II) which can facilitate the production of HO˙ (reaction (6)).

HO˙ is a highly active radical in the process of degradation of pollutants. The hydrocarbons RH was reduced to R˙ through dehydrogenation of HO˙. R˙ was then reduced to R⁺ by Fe³⁺-oxidation (reactions (20) and (21)). In addition, RH can be adsorbed on the catalyst surface (reaction (22)) and (reaction (23)) can occurred on the catalyst surface.

RH + HO˙/˙O₂⁻ → H₂O + R˙ (20)

R˙ + Fe³⁺-oxidation → R⁺ + Fe²⁺ (21)

RH + Ce⁴⁺ → RH–Ce⁴⁺ (22)

RH–Ce⁴⁺ → Ce³⁺ + R˙ + H⁺ (23)

The possible mechanisms can be illustrated in Fig. 11.

Fig. 11 Possible mechanisms of pollutants degradation in Fenton-like process using Ce/Fe as catalyst.

4. Conclusions

Heterogeneous Fenton-like process using Ce-doped Fe⁰ nano-composite as catalysts was an effective method for degradation of sulfamethazine. The Ce/Fe composite was more effective than Fe⁰ for degradation of SMT, especially at neutral pH. The degradation conditions were optimized as follows: pH = 6, T = 30 °C, catalyst dosage = 0.5 g L⁻¹, and H₂O₂ concentration = 12 mM, the removal efficiency of SMT was 100% at the optimal condition in 5 min. TOC removal efficiency was 62.2% in 60 min. The SMT degradation process followed the first-order degradation kinetic. Some small molecules organic acid and inorganic ions were detected during SMT degradation process and the possible degradation pathway was proposed. Ce-Doped zero valent iron can be an effective catalyst for Fenton-like degradation of SMT at neutral pH value.

Acknowledgements

The research was supported by the National Natural Science Foundation of China (51338005) and the Program for Changjiang Scholars and Innovative Research Team in University (IRT-13026).

References

1. J. Wang and S. Z. Wang, J. Environ. Manage., 2016, 182, 620–640.
2. I. Koyuncu, O. A. Arikan, M. R. Wiesner and C. Rice, J. Membr. Sci., 2008, 309, 94–101.
3. E. K. Putra, R. Pranowo, J. Sunarso, N. Indraswati and S. Ismadji, Water Res., 2009, 43, 2419–2430.
4. J. L. Wang and L. J. Xu, Crit. Rev. Environ. Sci. Technol., 2012, 42, 251–325.
5. N. De la Cruz, L. Esquius, D. Grandjean, A. Magnet, A. Tungler, L. F. De Alencastro and C. Pulgarín, Water Res., 2013, 47, 5836–5845.
6. L. Prieto-Rodríguez, I. Oller, N. Klamerth, A. Agüera, E. M. Rodríguez and S. Malato, Water Res., 2013, 47, 1521–1528.
7. A. El-Ghenymy, R. M. Rodríguez, C. Arias, F. Centellas, J. A. Garrido, P. L. Cabot and E. Brillas, J. Electroanal. Chem., 2013, 701, 7–13.
8. J. L. Wang and L. B. Chu, Radiat. Phys. Chem., 2016, 125, 56–64.
9. C. Orbeci, I. Untea, G. Nechifor, A. E. Segneanu and M. E. Craciun, Sep. Purif. Technol., 2014, 122, 290–296.
10. X. L. Zou, T. Zhou, J. Mao and X. H. Wu, Chem. Eng. J., 2014, 257, 36–44.
11. A. Ryu, S. Jeong, A. Jang and H. Choi, Appl. Catal., B, 2011, 105, 128–135.
12. A. Li, C. Tai, Z. S. Zhao, Y. W. Wang, Q. H. Zhang, G. B. Jiang and J. T. Hu, Environ. Sci. Technol., 2007, 41, 6841–6846.
13. X. Y. Wang, C. Chen, H. L. Liu and J. Ma, Water Res., 2008, 42, 4656–4664.
14. T. H. Zheng, J. J. Zhan, J. B. He, C. Day, Y. F. Lu, G. L. McPherson, G. Piringer and V. T. John, Environ. Sci. Technol., 2008, 42, 4494–4499.
15. Y. Zhang, M. Yang, X. M. Dou, H. He and D. S. Wang, Environ. Sci. Technol., 2005, 39, 7246–7253.
16. E. G. Heckert, S. Seal and W. T. Self, Environ. Sci. Technol., 2008, 42, 5014–5019.
17. L. J. Xu and J. L. Wang, Environ. Sci. Technol., 2012, 46, 10145–10153.
18. H. L. Li, C. Y. Wu, Y. Li and J. Y. Zhang, Environ. Sci. Technol., 2011, 45, 7394–7400.
19. M. Danilczuk, S. Schlick and F. D. Coms, Macromolecules, 2009, 42, 8943–8949.
20. L. Gubler and W. H. Koppenol, J. Electrochem. Soc., 2011, 159, B211–B218.
21. L. J. Xu and J. L. Wang, Appl. Catal., B, 2013, 142, 396–405.
22. Q. Ling, M. Yang, C. S. Li and A. M. Zhang, RSC Adv., 2014, 4, 4020–4027.
23. L. J. Xu and J. L. Wang, J. Hazard. Mater., 2011, 186, 256–264.
24. Y. H. Ni, X. W. Ge, Z. C. Zhang and Q. Ye, Chem. Mater., 2002, 14, 1048–1052.
25. W. L. Yan, A. A. Herzing, X. Q. Li, C. J. Kiely and W. X. Zhang, Environ. Sci. Technol., 2010, 44, 4288–4294.
26. A. P. Grosvenor, B. A. Kobe, M. C. Biesinger and N. S. McIntyre, Surf. Interface Anal., 2004, 36, 1564–1574.
27. P. Burroughs, A. Hamnett, A. F. Orchard and G. Thornton, J. Chem. Soc., Dalton Trans., 1976, 17, 1686–1698.
28. S. X. Yang, W. P. Zhu, Z. P. Jiang, Z. X. Chen and J. B. Wang, Appl. Surf. Sci., 2006, 252, 8499–8505.
29. J. D. Russell, Clay Miner., 1979, 14, 109.
30. B. Z. Tian and H. X. Tang, Environ. Chem., 1990, 9, 70–76.
31. V. F. Zolin, V. A. Kudryashova, V. V. Kuznetsova and T. I. Razvina, J. Appl. Spectrosc., 1974, 21, 1476–1479.
32. A. Turković, P. Dubček and S. Bernstorff, Mater. Sci. Eng., B, 1999, 58, 263–269.
33. T. Ohtsuka, K. Kubo and N. Sato, Corrosion, 1986, 42, 476–481.
34. B. M. Reddy, A. Khan, Y. Yamada, T. Kobayashi, S. Loridant and J. Volta, J. Phys. Chem. B, 2003, 107, 5162–5167.
35. X. M. Lin, L. P. Li, G. S. Li and W. H. Su, Mater. Chem. Phys., 2001, 69, 236–240.
36. G. S. Nolas, V. G. Tsoukala, S. K. Gayen and G. A. Slack, Phys. Rev. B: Condens. Matter Mater. Phys., 1994, 50, 150.
37. H. Jabeen, V. Chandra, S. Jung, J. W. Lee, K. S. Kim and S. B. Kim, Nanoscale, 2011, 3, 3583–3585.
38. Z. Q. Fang, J. H. Chen, X. H. Qiu, X. Q. Qiu, W. Cheng and L. C. Zhu, Desalination, 2011, 268, 60–67.
39. J. H. Chen, X. Q. Qiu, Z. Q. Fang, M. Yang, T. Pokeung, F. L. Gu, W. Cheng and B. Y. Lan, Chem. Eng. J., 2012, 181, 113–119.
40. S. X. Zha, Y. Cheng, Y. Gao, Z. L. Chen, M. Megharaj and R. Naidu, Chem. Eng. J., 2014, 255, 141–148.
41. J. Ma, M. X. Yang, F. Yu and J. H. Chen, J. Colloid Interface Sci., 2015, 444, 24–32.
42. Y. K. Liu, J. Hu and J. L. Wang, Radiat. Phys. Chem., 2014, 96, 81–87.
43. Z. Wan and J. L. Wang, Environ. Sci. Pollut. Res., 2016, 18542–18551.
44. Z. Wan and J. L. Wang, J. Chem. Technol. Biotechnol., 2016 DOI:10.1003/jctb.5072.
45. Z. Wan, J. Hu and J. L. Wang, J. Environ. Manage., 2016, 182, 284–291.
46. Z. Wan and J. L. Wang, Environ. Sci. Pollut. Res., 2016 DOI:10.1007/s11356-016-7768-9.
47. F. C. Moreira, R. A. R. Boaventura, E. Brillas and V. J. P. Vilar, Appl. Catal., B, 2017, 202, 217–261.
48. H. S. Ghandi, A. G. Piken, M. Shelef and R. G. Deloch, SAE Tech. Pap. Ser., 1976, 760201.
49. B. H. Bielski, D. E. Cabelli, R. L. Arudi and A. B. Ross, J. Phys. Chem., 1985, 14, 1041–1100.

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Footnote

† Electronic supplementary information (ESI) available. See DOI: 10.1039/c6ra23709f

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