CuO-Co3O4@CeO2 as a heterogeneous catalyst for efficient degradation of 2,4-dichlorophenoxyacetic acid by peroxymonosulfate

Journal of Hazardous Materials 381 (2020) 121209

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CuO-Co3O4@CeO2 as a heterogeneous catalyst for efficient degradation of 2, 4-dichlorophenoxyacetic acid by peroxymonosulfate ⁎

Wei Lia,b, Yuxin Lic, Deyun Zhangc, Yeqing Lanc, , Jing Guoc,

T

⁎

a

College of Resources and Environmental Sciences, Nanjing Agricultural University, Nanjing, 210095, PR China Jiangsu Tobacco Industrial Limited Company, Nanjing, 210011, PR China c College of Sciences, Nanjing Agricultural University, Nanjing, 210095, PR China b

G R A P H I C A L A B S T R A C T

A R T I C LE I N FO

A B S T R A C T

Editor: Danmeng Shuai

CuO-Co3O4@CeO2 nanoparticles used as a heterogeneous catalyst were prepared via a sol-gel method and characterized by various techniques. For comparison, a series of oxides was investigated for activating peroxymonosulfate (PMS) during the degradation of 2,4-dichlorophenoxyacetic acid (2,4-D). The results indicated that CuO-Co3O4@CeO2 exhibited the highest catalytic performance among the catalysts. Complete degradation of 2,4-D (20 mg/L) was realized within 45 min at 1 mM PMS, CuO-Co3O4@CeO2 loading of 0.07 g/L, and pH of 6. Recycling experiments confirmed that CuO-Co3O4@CeO2 was very stable, and the 2,4-D degradation efficiencies ranged from 100% to 97.5%, decreasing by only 2.5% after the fifth run. The outstanding catalysis of CuO-Co3O4@CeO2 resulted from the synergy of cerium, cobalt, and copper. Electron paramagnetic resonance and radical scavenger experiments confirmed the production of SO4• − and •OH radicals in the CuOCo3O4@CeO2/PMS system, which were responsible for efficient decomposition of 2,4-D. Furthermore, the combination of CuO-Co3O4@CeO2 andPMS was applied to treat natural water containing 2,4-D, and a high 2,4-D removal rate was also achieved. Based on these results, it was deduced that CuO-Co3O4@CeO2 can be utilized as a catalyst to activate PMS and destroy organic contaminants in aqueous solution.

Keywords: CuO-Co3O4@CeO2 nanoparticles Peroxymonosulfate Sulfate radicals 2,4-dichlorophenoxyacetic acid Degradation mechanism

⁎

Corresponding authors. E-mail addresses: [email protected] (Y. Lan), [email protected] (J. Guo).

https://doi.org/10.1016/j.jhazmat.2019.121209 Received 19 July 2019; Received in revised form 26 August 2019; Accepted 10 September 2019 Available online 11 September 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.

Journal of Hazardous Materials 381 (2020) 121209

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1. Introduction

method. Subsequently, the morphology, chemical composition, and elemental valence of the catalyst were characterized by X-ray powder diffraction (XRD), transmission electron microscopy (TEM), high-resolution transmission electron microscopy (HRTEM), and X-ray photoelectron spectroscopy (XPS). The catalytic performance of CuOCo3O4@CeO2 for activating PMS during degradation of 2,4-D was assessed. The effects of the main parameters, including PMS concentration, catalyst dosage, initial pH, inorganic anions, and water matrices on the destruction of 2,4-D were further assessed in detail. Furthermore, the stability of the catalyst and the possible mechanism and pathways of 2,4-D degradation were also explored.

2,4-Dichlorophenoxyacetic acid (2,4-D) has been used for many years as a systemic herbicide to selectively kill most broadleaf weeds (Brillas et al., 2000). 2,4-D can be found in commercial lawn herbicide mixtures, which often contain other active ingredients including mecoprop and dicamba. More than 1500 herbicide products contain 2,4-D as an active ingredient (Jaafarzadeh et al., 2017a). However, 2,4-D is a persistent and toxic substance that exerts potential destructive effects on humans and animals (Mantilla et al., 2009; Seck et al., 2012). The presence of 2,4-D may contaminate groundwater and soil because of its high water solubility and relatively poor biodegradability (Golshan et al., 2018; Chen et al., 2017; Aydın et al., 2005; Del Ángel-Sanchez et al., 2013). Therefore, it is necessary to develop an effective and economic method to remove 2,4-D from wastewater. In the past few years, several methods have been developed to remove 2,4-D from wastewater; specifically, adsorption and catalytic degradation (Kamaraj et al., 2014; Kearns et al., 2014; Kermani et al., 2018; Rodríguez et al., 2013). Adsorption is a feasible approach to remove 2,4-D due to its low cost and simple operation. However, this method does not destroy the structure of the 2,4-D molecule and may result in a large amount of sludge (Kamaraj et al., 2014; Kearns et al., 2014). In addition, the waste residues could cause secondary pollution in the environment. Advanced oxidation processes (AOPs) based on the generation of hydroxyl radicals (•OH), superoxide radicals (O2•−), and sulfate radicals (SO4• −) are considered more efficient approaches to remove persistent organic pollutants (Kermani et al., 2018; Rodríguez et al., 2013; Cai et al., 2018; Jaafarzadeh et al., 2017b). Notably, sulfate radicals have been widely investigated due to their high oxidation potential (E0 = 2.5–3.1 V) and superior selectivity to attack electro-rich pollutants compared with other free radicals (Tsitonaki et al., 2010; Li et al., 2018). Persulfate (PS) (Chen et al., 2018a) and peroxymonosulfate (PMS) (Deng et al., 2017a) can be activated by several methods to generate sulfate radicals. PMS is relatively easier to activate because of its asymmetrical structure (Oh et al., 2016a). It has been reported that ultrasound, ultraviolet light, heat, and transition metals can be utilized to activate PMS to generate sulfate radicals (Su et al., 2012; Guan et al., 2011; Yang et al., 2010; Anipsitakis and Dionysiou, 2003). Among these processes, transition metals have drawn more attention because of their low cost and simple operation. Transition metals, including Cu, Fe, Mn, and Co have been demonstrated to effectively activate PMS (Oh et al., 2016b; Saputra et al., 2016). Catalysts containing Co(II) have been reported to be more efficient for activating PMS than other transition ions, such as Mn(III), Cu(II), Fe(II), and Fe (III) (Bandala et al., 2007; Anipsitakis et al., 2005). Homogeneous catalysts exhibit some shortcomings compared to heterogeneous catalysts, such as high pH-dependence, high concentration of metal ions, and difficulty recycling the catalyst (Liang et al., 2012). Consequently, heterogeneous catalysts containing transition metals have been broadly explored to activate PMS. For instance, some monometallic oxides have been applied to activate PMS and degrade organic pollutants by oxidation (Anipsitakis and Dionysiou, 2003; Saputra et al., 2016). Nevertheless, monometallic oxides usually display lower catalytic performance than bimetallic oxides. It has been reported that bimetallic oxides containing cobalt and copper demonstrate high catalytic activity in inducing PMS to generate reactive oxygen species because of synergetic coupling of the active sites (Guan et al., 2013; Feng et al., 2015). Cerium (Ce) possesses a large oxygen storage capacity depending on the redox couple Ce(IV)/Ce(III), making more oxygen available for oxidation processes (Tian et al., 2019). 2,4-Dichlorophenol can be effectively degraded by PMS over Ce substituted 3D Mn2O3, and Ce in situ substitution results in excellent reductive capability, which is significant to generate free radicals (Tian et al., 2019). Based on this background, in this study, a novel nano-scale catalyst (CuO-Co3O4@CeO2) was prepared to activate PMS via a simple sol-gel

2. Materials and methods 2.1. Materials 2,4-D (> 98.8%) was obtained from Lideshi Chemical Technology Co., Ltd. (Beijing, China). Potassium peroxymonosulfate (KHSO5, 0.5 KHSO4, and 0.5 K2SO4) and sodium thiosulfate (Na2S2O3) were provided by Aldrich Inc. (St Louis, MO, USA). NaOH, H2SO4 and HNO3 were purchased from Kermel Chemical Reagent Co., Ltd. (Tianjin, China). Ce(NO3)3·6H2O, Cu(NO3)2·3H2O, and Co(NO3)2·6H2O were provided by Sinopharm Chemical Reagent Co., Ltd (Shanghai, China). Acetonitrile and formic acid used for the high performance liquid chromatography (HPLC) analysis were obtained from Tedia Co. (Fairfield, OH, USA). NaNO3, NaHCO3, Na2SO4, and NaCl were purchased from Xilong Chemical Co., Ltd (Shantou, China). Methanol (MeOH) and tert-butyl alcohol (TBA) were supplied by Nanjing Chemical Reagent Co., Ltd. (Nanjing, China). Furfuryl alcohol (FFA) and p-benzoquinone (BQ, 99%) were purchased from Macklin Biochemical Co., Ltd (Shanghai, China) and Aladdin Industry Corp. (Shanghai, China), respectively. The two real water samples were collected from the Yangtze River and Xuanwu Lake (Nanjing, China), respectively. Other chemicals used in this experiment were of analytical grade. 2.2. Preparation of the catalysts The CuO-Co3O4@CeO2 nanoparticles were synthesized using the classic sol-gel preparation method (Ding et al., 2013). First, 0.01 mol of Ce(NO3)3·6H2O, 0.005 mol Co(NO3)2·6H2O, and 0.005 mol Cu (NO3)2·3H2O were dissolved in 200 mL deionized water to form a mixed salt solution. Then, 0.02 mol of citric acid was introduced to form citric acid chelate compounds under vigorous mechanical stirring for 10 h at 80 °C. The obtained viscous gel was dried at 100 °C for 18 h. Finally, the resin was ground into a fine powder and calcined at 600 °C for 5 h to form the CuO-Co3O4@CeO2 catalyst. The sample was repeatedly washed to neutral pH with pure water and methanol and dried overnight at 60 °C. In addition, the bimetallic oxides (CuO@CeO2, Co3O4@CeO2, and CuO@Co3O4) and the monometallic oxides (CeO2, CuO, and Co3O4) were synthesized using the same sol-gel method described above. 2.3. Characterization The XRD patterns of the catalysts were collected by a D8 Advance Xray diffractometer (Bruker, Hamburg, Germany) with a Cu-Kα radiation source of 40 kV and 35 mA. The morphological studies were conducted using TEM and HRTEM (JEM-2010FEF, Jeol, Tokyo, Japan). The Brunauer–Emmet–Teller (BET) specific surface areas of the catalysts were measured by N2 adsorption isotherms at 77 K using a surface area and porosity analyzer (ASAP 2460; Micromeritics, Norcross, GA, USA). The pore size distribution plots and pore volume were obtained using the Barrett–Joyner–Halenda method. XPS was analyzed using an X-ray photoelectron spectrometer (ESCALAB250i, Thermo Fisher Scientific, Waltham, MA, USA). 2

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2.4. Experimental procedure 2,4-D degradation was assessed in a glass flask (250 mL) containing 100 mL of 2,4-D (20 mg/L) and the desired dosage of PMS. The initial solution pH was adjusted to the desired pH (2–12) with diluted H2SO4 and NaOH solutions. The change in the suspension volume was negligible due to the small volume of the H2SO4 and NaOH solutions added. Then, the flask was fixed in a reciprocating shaker (COS-110 × 30, Shanghai Bilon, Shanghai, China) for 10 min at 180 rpm and 25 °C. Next, the desired amount of catalyst (0.01–0.1 g/L) was introduced into the flask to initiate the reaction. At set intervals, 2 mL of the suspension was extracted with a plastic syringe and immediately filtered through a 0.22 μm membrane filter into a clear tube. The collected solution (1 mL) was mixed with 1 mL Na2S2O3 (10 mM) to stop the radical reaction. Finally, the solution was analyzed for the concentration of remaining 2,4-D (Xu et al., 2018). In addition, the solution after the reaction was collected from a LC-18 solid phase extraction column (Supelclean; Supleco, Bellefonte, PA, USA), and the extract was eluted with 1 mL methanol to determine the intermediates.

Fig. 1. X-ray diffraction pattern of the synthesized CuO-Co3O4@CeO2.

2.5. Analytical methods

diffraction peaks attributed to CuO (JCPDS NO. 45-0937), Co3O4 (JCPDS NO. 42-1467), and CeO2 (JCPDS NO. 43-1002) were clearly observed in the CuO-Co3O4@CeO2 catalyst, suggesting that cobalt and copper were successfully anchored on cerium oxide. Compared to the pure monometallic oxides, all CuO-Co3O4@CeO2 diffraction peaks appeared shorter and broader, suggesting a smaller average crystallite size and a distinct relationship of mutual influence (Xu et al., 2018). Moreover, the XRD spectra of the bimetallic oxides (CuO@CeO2, Co3O4@CeO2, and CuO@Co3O4) and the monometallic oxides (CuO, Co3O4, and CeO2) are illustrated in Fig. S1(a) and (b), respectively, indicating that these oxides were also successfully synthesized. An N2 sorption-desorption analysis was carried out to determine the textural properties of the catalysts, including the ternary metallic and bimetallic oxides. The N2 adsorption-desorption isotherms for the catalysts are illustrated in Fig. S2 and the BET surface area and the pore structural parameters of the catalysts are summarized in Table S1. The BET surface areas of CuO-Co3O4@CeO2, CuO@CeO2, Co3O4@CeO2, and CuO@Co3O4 were 73.03, 38.22, 35.42, and 14.34 m2/g, respectively. Correspondingly, their pore volumes were 0.2369, 0.0908, 0.0649, and 0.0387 cm3/g, respectively. These results indicate that CuO-Co3O4@CeO2 exhibited the largest surface area and the largest pore volume among the catalysts, which is in favor of providing more catalytic active sites. The detailed morphology of CuO-Co3O4@CeO2 was obtained from the TEM and HRTEM analyses. As shown in Fig. 2a, the TEM image of the CuO-Co3O4@CeO2 demonstrated an average particle size of 5–20 nm. The reflections with d-spacing values of 0.312 and 0.241 nm (Fig. 2b) resulted from the CeO2 cubic crystal facet (111) and Co3O4 facet (311), respectively. Moreover, selected area electron diffraction was conducted to authenticate the structure of the CuO-Co3O4@CeO2 composite. In Fig. 2c, the bright electron diffraction rings demonstrated the formation of a polycrystalline structure, which was consistent with the XRD results. The energy dispersive X-ray spectrum of the CuOCo3O4@CeO2 composite is illustrated in Fig. 2d, and the atomic percent ratio of the elements was in fair agreement with the dosage of the elements used during preparation of the catalyst. The elemental mapping images are displayed in Fig. 2e, indicating that Ce, Cu, Co, and O were uniformly distributed in the CuO-Co3O4@CeO2 catalyst and it was not a random mixture of the corresponding three monometallic oxides (Yao et al., 2017; Somekawa et al., 2011; Khan et al., 2017). The surface elemental compositions and chemical valences of the CuO-Co3O4@CeO2 were analyzed using the XPS spectra. As shown in Fig. 3a, the Ce 3d XPS spectrum of CuO-Co3O4@CeO2 before and after the reaction was deconvoluted into eight peaks assigned to Ce(III) and Ce(IV), corresponding to Ce 3d3/2 and Ce 3d5/2, respectively. This result

2,4-D concentration was measured by HPLC (1260, Agilent Technologies, Palo Alto, CA, USA) with an Eclipse plus C18 column (250 mm ×4.6 nm, 5 μm) and a diode-array-detector at a wavelength of 280 nm. The mobile phase was acetonitrile/water (with 0.2% formic acid) at a volume ratio of 40:60 and a flow rate of 1.2 mL min−1. A 10μL sample was automatically injected while the column temperature was set to 30 °C. Solution pH values were monitored on a CyberScan pH2100 Bench Meter (Eutech Instruments, Vernon Hills, IL, USA) after three-point calibration. The total organic carbon (TOC) levels of the 2,4-D solution before and after the reaction and samples of deionized water, Yangtze River, and Xuanwu Lake water were measured with a Shimadzu TOC-L analyzer (Shimadzu, Tokyo, Japan). The concentrations of the metal ions (cerium, cobalt, and copper) dissolved from the catalyst were measured by atomic absorption spectrometry (AA-7020, Shimadzu). The intermediates of 2,4-D degradation were analyzed by gas chromatograph/mass spectrometry (GC–MS) (Thermo Scientific ISQ, Waltham, MA, USA). Separation and measurement were achieved with a DB-17 (30 m ×0.25 mm, 0.25 μm) capillary column with high purity (99.99%) helium as the carrier gas at a flow rate of 1.0 mL/min. The inlet temperature was set to 250 °C with an injection volume of l.0 μL. The initial column temperature was set to 35 °C for 1 min, and then increased to 180 °C at a rate of 7.0 °C/min and maintained constant for 1 min. Next, the oven temperature was increased to 240 °C at a rate of 10 °C/min and stabilized for 1 min. The concentration of chloride ions (Cl−) released from the reaction and the anions, including Cl−, NO3−, HCO3−, and SO42− in the different water matrices were analyzed by ion chromatography (ICS-600, Thermo Scientific). UV254 was measured by a UV–vis spectrophotometer (Alpha-1502, Puyuan, China). Electron paramagnetic resonance (EPR) experiments were conducted on a Bruker EMX 10/12 spectrometer using 5,5-dimethyl-1pyrroline N-oxide (DMPO) to capture the radical signals. Microwave frequency was 9.773 GHz. The measurement at the center of the field was 3480 G. Microwave power was 19.45 mW and the modulation amplitude was 2 Gs. The electrochemical measurements were performed on an electrochemical workstation (CHI660e, CH Instruments, Shanghai, China) with a three-electrode cell at room temperature. The details of the processes are described in text S1. 3. Results and discussion 3.1. Characterization of the catalysts The XRD pattern of CuO-Co3O4@CeO2 is displayed in Fig. 1. The 3

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Fig. 2. (a) Transmission electron microscopic (TEM) image, (b) high resolution TEM image, (c) selected area electron diffraction image, (d) energy-dispersive image, and (e) elemental mapping images of cerium, cobalt, copper, and oxygen of the synthesized CuO-Co3O4@CeO2.

different reaction systems were compared and the results are illustrated in Fig. 4. Only 8.3% of the 2,4-D (20 mg/L) was degraded by PMS (1 mM) alone within 45 min, suggesting that the oxidative degradation of 2,4-D by PMS was weak. Less than 3.2% of the 2,4-D was removed in the presence of CuO-Co3O4@CeO2, suggesting that adsorption of 2,4-D by the catalyst was negligible. However, almost 100% of the 2,4-D was degraded by a combination of CuO-Co3O4@CeO2 and PMS within 45 min, demonstrating that CuO-Co3O4@CeO2 possesses high catalytic performance for activating PMS during 2,4-D degradation. Only 9.8%, 10.1%, 12.2%, 28.3%, 42.1%, and 53.4% of the 2,4-D were removed in the CeO2/PMS, CuO/PMS, CuO@CeO2/PMS, Co3O4/PMS, CuO@Co3O4/PMS, and Co3O4@CeO2/PMS systems, respectively. These results demonstrate that the catalytic activity of cobalt oxide was superior to that of copper oxide or cerium oxide and that the catalytic performance of the oxides was in the order: monometallic oxides < bimetallic oxides < CuO-Co3O4@CeO2. In this study, a specific experiment was designed to investigate the homogeneous catalytic role of the metal ions dissolved from the catalyst during degradation of 2,4-D by PMS. First, 2,4-D was mixed with CuOCo3O4@CeO2, and the mixture was shaken at 25 °C for 45 min. Then, the suspension was equally divided into two parts, and the catalyst was removed from one of the parts. Finally, PMS was introduced into the two solutions to initiate the reaction again. Consequently, only 14.1% of the 2,4-D was further degraded within 45 min in the reaction system where the catalyst was removed. However, the residual 2,4-D was

implies that Ce exists in the catalyst as the Ce(III) and Ce(IV) species (Tian et al., 2019). The Co 2p XPS spectra of the CuO-Co3O4@CeO2 before and after the reaction are illustrated in Fig. 3b. Before the reaction, two main peaks appeared at 779.8 and 794.4 eV, corresponding to Co 2p3/2 and Co 2p1/2, respectively. The two Co 2p3/2 peaks at 779.3 and 780.8 eV were attributed to Co(II) and Co(III), respectively, whose atomic ratio is 24.8:75.2 (Liu et al., 2018; Hu et al., 2017). Fig. 3c shows the Cu 2p XPS spectra of CuO-Co3O4@CeO2 before and after the reaction. Combined with the XRD pattern, the peak at a binding energy of 934.1 eV along with the satellite peak was ascribed to the 2p3/2 region of Cu(II), indicating that Cu existed only as a Cu(II) species in the catalyst (Ding et al., 2013). The O1 s spectra in Fig. 3d indicates the existence of two surface oxygen species in the catalyst. The peak at the binding energy of 529.6 eV was assigned to lattice oxygen (OL) bonding to metal cations, and the peak around 531.7 eV was ascribed to the surface hydroxyl species or the adsorbed oxygen (OH) (Liu et al., 2018). The proportion of lattice oxygen decreased from 55.2% before the reaction to 52.1% after the reaction. In contrast, OH increased from 44.8% before the reaction to 47.9% after the reaction, suggesting that the lattice oxygen was oxidized during the reaction (Deng et al., 2017a). 3.2. Degradation of 2,4-D in different systems To assess the catalytic performance of the prepared CuOCo3O4@CeO2 for activating PMS, the 2,4-D degradation rates in 4

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Fig. 3. X-ray photoelectron spectroscopy spectra of Ce 3d (a), Co 2p (b), Cu 2p (c) and O 1s (d) of CuO-Co3O4@CeO2 before and after the reaction.

by a combination of Fe(III) and H2O2, a greater amount of oxidant and a lower pH were needed. This finding further verifies that CuOCo3O4@CeO2 is a promising catalyst for 2,4-D degradation by PMS. 3.3. Factors impacting 2,4-D degradation by CuO-Co3O4@CeO2/PMS 3.3.1. Effect of catalyst loading The effect of catalyst loading on 2,4-D degradation by PMS was investigated in a series of batch experiments, and the results are shown in Fig. 5a. The degradation rate of 2,4-D increased from 66.7% to 100% within 45 min as the amount of catalyst was increased from 0.01 to 0.07 g/L. To further compare the 2,4-D degradation rates by PMS at different catalyst loadings, the logarithmic plots of the 2,4-D concentration (c/c0) versus reaction time (t) were drawn, and some parameters are listed in Table S2. The rate constant (k) of 2,4-D degradation increased from 0.0524 min−1 at 0.01 g/L catalyst to 0.1344 min−1 at 0.07 g/L catalyst. This is because more catalyst provides more active sites for activating the PMS, which leads to the production of more reactive oxygen species and more efficient 2,4-D degradation. Nevertheless, the degradation efficiency of 2,4-D began to slightly decline with a further increase of the catalyst dosage (> 0.07 g/L). Correspondingly, the rate constant (k) decreased from 0.1344 to 0.1098 min-1 as the catalyst was increased from 0.07 to 0.1 g/L. These results suggest that the optimal catalyst dosage for degrading 2,4-D (20 mg/L) by PMS (1 mM) is 0.07 g/L at the initial pH of 6.

Fig. 4. Degradation of 2,4-D in different systems. Experimental conditions: [2,4-D] =20 mg/L, [PMS] =1 mM, [catalyst] =0.07 g/L, pH = 6, T =25 °C.

almost completely decomposed within 30 min in the other reaction system where the catalyst was not removed (see Fig. S3). These results confirm that the contribution of dissolved cobalt ions to the degradation of 2,4-D was not important in the CuO-Co3O4@CeO2/PMS system. Therefore, it is concluded that the rapid removal of 2,4-D by PMS is ascribed to the excellent catalysis of CuO-Co3O4@CeO2. Furthermore, to further assess the catalytic performance of CuOCo3O4@CeO2, the degradation efficiencies of 2,4-D from other AOP techniques reported in previous studies were used to make a comparison (see Table 1). 2,4-D was also efficiently degraded in the TiO2@CuFe2O4/UV/PMS, nano-Fe2O3/PMS, heat/PS, electricity/PMS and FeS/PS systems. Nevertheless, the 2,4-D degradation efficiencies from these methods were far lower than that obtained in the CuOCo3O4@CeO2/PMS system. Although 2,4-D was more rapidly removed

3.3.2. Effect of PMS concentration The effect of PMS concentration (0.1–1.2 mM) on 2,4-D degradation was investigated, and the results are depicted in Fig. 5b. 2,4-D degradation efficiency increased with an increase in the PMS concentration from 0.1 to 1.0 mM. Accordingly, the rate constant (k) listed in Table S2 rose from 0.0403 to 0.1344 min−1, which was ascribed to the fact that the higher PMS dosage generated more radicals. Nevertheless, the 2,4-D degradation rate was not further enhanced when PMS concentration was increased from 1.0 to 1.2 mM. This result may be 5

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Table 1 Comparison of the 2,4-D degradation in the system CuO-Co3O4@CeO2/PMS with those from previous studies. 2,4-D concentrations (mg/L)

Catalyst

Oxidant

Initial pH

Efficiency

Ref.

20 10 40 22 100 20 20

TiO2@CuFe2O4/UV, 0.2 g/L FeS, 0.15 g/L current density, 40 mA/cm2 Fe3+, 0.1 mM heat energy nano-Fe2O3, 0.5 g/L CuO-Co3O4@CeO2, 0.07 g/L

PMS, 0.3 mM PS, 1.25 mM PMS, 2.0 mM H2O2, 10.0 mM PS, 18 mM PMS, 3 mM PMS, 1 mM

6.5 4.5 4 2.8 3.5 6.0 6.0

94.9%, 60 min 100%, 120 min 83.1%, 90 min 100%, 40 min 100%, 120 min 52.3%, 60 min 100%, 45 min

(Golshan et al., 2018) (Chen et al., 2017) (Jaafarzadeh et al., 2018) (Lee et al., 2003) (Cai et al., 2018) (Jaafarzadeh et al., 2017b) Present study

explained by the fact that excess SO4• − reacted with PMS and itself according to Eqs. (1) and (2) (Li et al., 2014). Consequently, 2,4-D degradation could not be further improved at a higher concentration of PMS (> 1 mM). Based on these results, 1.0 mM was accepted the optimal PMS concentration in the following 2,4-D degradation experiments. HSO5− + SO4•− → SO5•− + SO42− + H+ SO4

•−

+

SO4•−

→ S2O8

2−

SO4

−

+ OH

HSO5−

↔

•

+H

+

(7)

+

•

+ OH + e

−

→ H2O

(8)

SO4•− + Cl− → SO42− + Cl• •

Cl + H2O → ClHO

• −

+H

(9)

+

(10)

The degradation efficiency of 2,4-D clearly decreased when HCO3− was introduced compared to the other anions. It has been reported that the presence of bicarbonate exerts a relatively strong effect on the degradation of bisphenol A (Liu et al., 2018). Thus, bicarbonate is usually used as a radical scavenger to react with SO4•−and •OH as described by (Eqs. (11) and (12)) (Guan et al., 2013; Ghauch et al., 2017). Therefore, bicarbonate suppressed 2,4-D degradation. SO4•− + HCO3− → SO42− +H+ + CO3•− •

OH +

HCO3−

→ H2O + CO3

•−

(11) (12)

3.3.5. Effect of different water matrices The degradation of 2,4-D by CuO-Co3O4@CeO2/PMS in different water matrices was further investigated, and the results are depicted in Fig. 7. Some main parameters of the water samples are listed in Table 2. The degradation efficiencies of 2,4-D in deionized water, Yangtze River water, and Xuanwu Lake water were 95.1%, 91.2%, and 84.6%, respectively. The decrease in 2,4-D degradation from the latter two water matrices compared with deionized water probably resulted from the existence of inorganic anions, natural organic matter, and synthetic organic compounds (Deng et al., 2017b). Organic compounds and some inorganic anions would compete with 2,4-D for free radicals, leading to a slight decrease in 2,4-D degradation. On the whole, the combination of CuO-Co3O4@CeO2 and PMS exhibited good performance for removing 2,4-D from natural water.

(3)

3.4. Determination of free radicals

2−

(4)

pka=9.4

(5)

Methanol (MeOH), as a scavenger of both •OH and SO4•−, and TBA, as a scavenger of •OH (Nie et al., 2014), were introduced into the

→ OH+SO4

SO52−

H+ + SO4•− + e− → HSO4−

3.3.4. Effect of different inorganic anions To investigate the effects of different anions on removal of 2,4-D, several inorganic anions with the same concentrations were introduced into the CuO-Co3O4@CeO2/PMS/2,4-D system. Fig. 6 shows that suppression by the anions was in the order of HCO3− > Cl− > SO42− ≈ NO3−. The presence of nitrate and sulfate ions slightly reduced 2,4-D degradation efficiency. This is because nitrate and sulfate ions seldom react with free radicals (Liu et al., 2018; Nie et al., 2014). The 2,4-D degradation rate decreased by 8.9% in the presence of Cl− (10 mM) compared with the control. The suppression of chloride ions was considered, as a portion of the sulfate radicals was consumed by Cl− to produce less reactive chlorine species, as described in Eqs. (9) and (10) (Golshan et al., 2018; Hu et al., 2018).

(1)

3.3.3. Effect of initial pH Solution pH is one of the most important parameters influencing the degradation of organic contaminants during catalytic oxidation processes. Therefore, the effect of initial pH ranging from 2 to 12 on 2,4-D degradation was further examined, and the results are shown in Fig. 5c. 2,4-D was completely removed within 45 min at the initial pH of 6.0. 2,4-D removal efficiency declined slightly at the initial pHs of 4.0 and 8.0. However, a significant decrease in degradation was observed when the initial solution pH was < 4 or > 8. The final pH after catalytic degradation was monitored to understand this phenomenon. As shown in Table S3, the final pH values were almost the same as the initial pH ranging from 4.0 to 8.0. This is because the CuO-Co3O4@CeO2 catalyst provided some buffering capacity (discussed later). The rate constants listed in Table S2 were also close. These results are in accordance with the reports of Hu et al. (2017) and Lei et al. (2015). In contrast, the rate constants decreased from 0.0948 min−1 at the initial pH of 8 to 0.0211 min-1 at the initial pH of 12 (see Table S2). This finding could be partially explained by the fact that the SO4• − radicals produced were transformed to •OH radicals under alkaline conditions according to Eqs. (3) and (4) (Tsitonaki et al., 2010), which did not favor degradation of 2,4-D due to the much shorter lifetime of the •OH radicals (< 1 μs) than that of SO4• − radicals (10−30 μs). It might also be possible that the interaction between OH− and the catalyst at a high pH affected the catalytic active sites. Furthermore, the surface charges of the metal oxide catalyst strongly depend on its pHpzc and solution pH (Schwarzenbach et al., 2005). The pHpzc of CuO-Co3O4@CeO2 was determined to be 6.74 (see Fig. S4). This means that the surface of the CuO-Co3O4@CeO2 is negatively charged (≡M-O−) at pHs > 6.74 (consume OH−) and carries a positive charge (≡M−OH2+) at pHs < 6.74 (consume H+). SO52− appears as the major species of HSO5− at pHs > 9.4 due to the pKa2 of PMS (9.4) (Eqs. (5) and (6)) (Wang et al., 2017). Hence, a high pH leads to electrostatic repulsion between the catalyst and PMS, resulting in a sharp decline in the 2,4-D degradation rate. Under an acidic condition of pH 2.0, H+ ions react with SO4•− radicals according to Eqs. (7) and (8) (Lai et al., 2018). In this case, the solution pH value was expected to rise, which was confirmed by the final pH (2.26). This result implies that the free radicals produced were partially consumed. Consequently, the 2,4-D degradation rate decreased by 83.4% compared to that at the initial pH of 6.

•−

(6)

H

(2)

SO4•− + H2O → •OH + H+ + SO42−

HSO5− + OH−→ SO52−+ H2O

6

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Fig. 6. Effect of different anions on 2,4-D degradation in the CuOCo3O4@CeO2/PMS system. Experimental conditions: [2,4-D] =20 mg/L, [PMS] =1 mM, [catalyst] =0.07 g/L, pH = 6, T =25 °C.

Fig. 7. Degradation of 2,4-D by CuO-Co3O4@CeO2/PMS in different matrices. Table 2 Parameters of the obtained water samples.

pH UV254 (cm−1) TOC(mg/L) Cl− (mM) SO42− (mM) NO3− (mM)

Deionized Water

Yangtze River

Xuanwu Lake

6.97 0.003 0.212 – – –

7.34 0.043 2.543 0.322 0.244 0.188

7.29 0.048 3.389 0.431 0.187 0.194

demonstrate that both SO4•− and •OH are crucial oxidants responsible for efficient decomposition of 2,4-D in the CuO-Co3O4@CeO2/PMS system. Besides SO4•− and •OH, 1O2 and O2• −are reactive oxygen species in the PMS activation processes (Wang et al., 2018; Zhang et al., 2014). In this study, FFA and BQ were used as 1O2 and O2• − scavengers, respectively. However, adding FFA and BQ at pH 6 had little effect on 2,4-D degradation (Fig. S5), suggesting that 1O2 and O2• − were not reactive oxygen species in the reaction system. The reported non-radical process in the CuO-Co3O4@CeO2/PMS system may be ruled out. These results demonstrate that different reactive oxygen species can be detected in different reaction systems in the presence of PMS. To further validate the active radicals produced by the CuOCo3O4@CeO2/PMS system, the EPR spectrum using DMPO as the

Fig. 5. Effect of catalyst dosage (a), PMS concentration (b) and initial pH (c) on the degradation of 2,4-D in the CuO-Co3O4@CeO2/PMS system. Experimental conditions: [2,4-D] =20 mg/L, [PMS] =1 mM, [catalyst] =0.07 g/L, pH = 6, T =25 °C.

reaction system to compare the contributions of different radicals to 2,4-D degradation. As shown in Fig. 8a, 2,4-D degradation efficiency decreased by 40% in the presence of 10 mM TBA compared with the control (no radical scavenger), indicating that •OH contributed substantially to 2,4-D degradation. In contrast, the degradation efficiency of 2,4-D decreased by 88% when MeOH was added. These results 7

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Fig. 9. Recycling of CuO-Co3O4@CeO2 during degradation of 2,4-D. Experimental conditions: [2,4-D] =20 mg/L, [PMS] =1 mM, [catalyst] =0.07 g/L, pH = 6, T =25 °C.

only 2.5% after five runs. The slight decline in 2,4-D degradation efficiency may have resulted from the interference of the intermediates adsorbed on the surface of the catalyst (Hammouda et al., 2017) and the weak dissolution of metal ions from the catalyst surface. To examine the dissolution of metal ions during the catalytic reaction, the concentrations of the three metal ions were monitored at different solution pHs and the results are illustrated in Fig. S6. The concentrations of the dissolved metal ions were very low (< 0.7 mg/L) even at pH 2, and further decreased as the solution pH was increased from 2 to 12. 3.6. Possible PMS activation mechanism PMS is activated by Co(II) (Ce(III)) to produce sulfate radicals (Eq. (13)) (Tian et al., 2019). It was noted from Fig. 3a and 3b that the ratios of Co(III) /Co(II) and Ce(IV)/Ce(III) declined slightly after the reaction, suggesting that Co(III) (Ce(IV)) was reduced to Co(II) (Ce(III)) during the reaction. The increase in Co(II) and Ce(III) ions may be attributed to the interaction between the catalyst and PMS (Eq. (14)) (Feng et al., 2015; Tian et al., 2019). As mentioned previously, lattice oxygen (OL) declined from 55.2% to 52.1%, and the surface hydroxyl species or adsorbed oxygen (OH) rose from 44.8% to 47.9% (see Fig. 3d), suggesting that the oxygen-containing groups participated in activating PMS. This inference was also supported by Chen et al. (2019) and Deng et al. (2017a). OL resulted in the reduction of both Ce(IV) and Co(III) ions, as described in Eq. (15) (Deng et al. (2017a)). A similar reaction (Eq. (16)) was expected to occur with Cu(II) (Ding et al., 2013). Subsequently, Co(III) and Ce(IV) were rapidly reduced to Co(II) and Ce(III) by the Cu(I) produced (Eq. (17)), which was converted to Cu(II). This may be why Cu(I) was not detected on the catalyst after the reaction. Hence, the synergy of cobalt, copper, and cerium ions maintained the high concentrations of Co(II) and Ce(III) ions in the catalyst, which efficiently activated PMS to generate free radicals according to Eq. (13). Electrochemical measurements were conducted on an electrochemical workstation to further reveal the catalysis of CuOCo3O4@CeO2 resulting from the synergy of cerium, cobalt, and copper. As depicted in Fig. S7a, CuO-Co3O4@CeO2 exhibited a higher current density than that of CuO@CeO2 or CuO@Co3O4. The EIS Nyquist plots in Fig. S7b indicated that the CuO-Co3O4@CeO2 had a lower resistance and a higher electric conductivity than CuO@CeO2 or CuO@Co3O4. These results suggest that greater and faster electron transfer improved the activation of PMS by CuO-Co3O4@CeO2. Nevertheless, the electrochemical properties of CuO-Co3O4@CeO2 and Co3O4@CeO2 were very similar, which did not agree with the results illustrated in Fig. 4. Fig. 4 shows that the catalytic performance of CuO-Co3O4@CeO2

Fig. 8. (a) Effect of different scavengers on 2,4-D degradation. Experimental conditions: [2,4-D] =20 mg/L, [PMS] =1 mM, [catalyst] =0.07 g/L, pH = 6, T =25 °C, [TBA] = [MeOH] =10 mM. (b) Electron paramagnetic resonance spectra of different systems during 2,4-D degradation. Experimental conditions: [PMS] =0.2 mM, [DMPO] =10 mM, [catalyst] =0.02 g/L, pH = 6, T =25 °C.

radical trap was analyzed, and the results are shown in Fig. 8b. Typical DMPO-SO4•−and DMPO-•OH signals were detected in the CuOCo3O4@CeO2/PMS/DMPO system, confirming again that PMS could be activated by CuO-Co3O4@CeO2 to generate •OH and SO4•− radicals. Nevertheless, the intensity of the characteristic peaks of the DMPOSO4•−adducts was much weaker than that of DMPO-•OH. This phenomenon was also observed by Yu et al.(Yue et al., 2018) and (Zhu et al. (2019) because DMPO-SO4•− adducts undergo rapid nucleophilic substitution by H2O or OH− to form DMPO-•OH adducts (Timmins et al., 1999). In contrast, •OH is derived from SO4•− and H2O or OH− reacts as described in Eqs. (3) and (4).

3.5. Stability of the catalyst Stability of the catalyst is an important property. To investigate the stability of the CuO-Co3O4@CeO2 nanoparticles as a heterogeneous catalyst, five consecutive runs of the catalyst to activate PMS during 2,4-D degradation were carried out, and the results are illustrated in Fig. 9. The catalyst exhibited good performance and stability over the five cycling experiments. The degradation efficiencies of 2,4-D ranged from 100% after the first run to 97.5% after the fifth run, decreasing by 8

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through free radical attack and removal of the para-positioned and ortho-positioned Cl−. Due to the electrophilic attack, 1,2,4-benzenetriol was generated and further transformed to succinic acid (Nagashree et al., 2018). In addition, 2,4-DCP was also attacked by free radicals to form 4,6-dichlororesorcinol (Chen et al., 2017). 2,4-DCP reacted with Cl• generated from Cl− to form 2,4,6-TCP. Then, 2,6-DCP was produced through dechlorination (Cai et al., 2018). Finally, the succinic acid and other intermediates were mineralized to H2O and CO2, which was demonstrated by the detected TOC and chloride ions in the solution after the reaction (see Fig. S9). The removal of TOC reached 44.5% at the end of the reaction. In addition, the dechlorination rate also increased simultaneously, and approximately 0.074 mM Cl− was detected after the degradation process. This finding means that 40.9% of the theoretical chlorine concentration in the 2,4-D molecules (0.181 mM) was released in solution, which was in accordance with the 2,4-D mineralization result (Chen et al., 2017).

during degradation of 2,4-D by PMS was better than that of Co3O4@CeO2. As discussed previously, this may be ascribed to the fact that CuO-Co3O4@CeO2 displayed a much larger surface area and a much larger pore volume than Co3O4@CeO2, which is conducive to providing more catalytic active sites to activate PMS. As a result, CuOCo3O4@CeO2 with excellent electron-conductive properties serves as an electron transfer mediator to accelerate decomposition of PMS. In addition, •OH was present in the reaction system by free scavenging and EPR and was generated according to Eqs. (3) and (4). Consequently, 2,4-D was effectively degraded by PMS in the presence of the CuO-Co3O4@CeO2 catalyst (Eq. (18)). ≡Co(II) /Ce(III) + HSO5− → ≡Co(III) /Ce(IV) + SO4•− + OH− (13) ≡Co(III) /Ce(IV) + HSO5− → ≡Co(II) /Ce(III)+ SO5•− + H+ ≡Co(III) /Ce(IV) + ≡Cu(II) + HSO5

−

O2−

→ ≡Co(II) /Ce(III) + O2

→ ≡Cu(I) +

SO5•−

+H

+

(14) (15) (16)

≡Co(III) /Ce(IV) + ≡Cu(I) → Co(II) /Ce(III) + ≡Cu(II)

(17)

4. Conclusions

SO4•− /•OH + 2,4-D → intermediates → CO2 + H2O + Cl−

(18)

CuO-Co3O4@CeO2 nanoparticles were successfully synthesized and were applied to activate PMS for the first time. The CuO-Co3O4@CeO2 nanoparticles as a heterogeneous catalyst exhibited excellent performance during the degradation of 2,4-D by PMS. Under the optimal conditions, 2,4-D (20 mg/L) was completely decomposed within 45 min. The high catalytic activity of CuO-Co3O4@CeO2 was ascribed to the synergy of cerium, cobalt, and copper ions on the surface of the catalyst. The EPR spectral analysis and radical quenching experiment revealed that SO4•−and •OH radicals were responsible for the rapid degradation of 2,4-D. The CuO-Co3O4@CeO2 catalyst was also highly stable during the catalytic reaction. No obvious decline in catalytic activity was observed after five recycling runs. Based on the GC–MS analyses, eight intermediates were determined and three possible pathways of 2,4-D degradation were proposed. Therefore, we deduced that CuO-Co3O4@CeO2 is a novel and promising catalyst for activating PMS during degradation of organic contaminants.

3.7. Possible 2,4-D degradation pathways by CuO-Co3O4@CeO2/PMS To identify the possible 2,4-D degradation pathways with CuOCo3O4@CeO2 and PMS, the intermediates were detected using the GC–MS data shown in Fig. S8. Seven main intermediates were identified, including 2,4-dichlorophenol (2,4-DCP), 2-chloro-1,4-hydroxyquinone (2-CHQ), 2,6-dichlorophenol (2,6-DCP), 2,4,6-trichlorophenol (2,4,6-TCP), 4-CHlorocatechol (4-CCA), 4,6dichlororesorcinol, and succinic acid. Based on the results obtained in this study, three possible 2,4-D degradation pathways are proposed (shown in Fig. 10). The CeO bond in the phenoxy group of 2,4-D was first attacked by the free radicals (•OH and SO4•−) and was converted to 2,4-DCP. Then, 2-CHQ and 4-CCA were produced from 2,4-DCP

Fig. 10. Possible 2,4-D degradation pathways in the CuO-Co3O4@CeO2/PMS system. 9

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Acknowledgment

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