Activation of persulfate by CuOx@Co-LDH: A novel heterogeneous system for contaminant degradation with broad pH window and controlled leaching

Chemical Engineering Journal 335 (2018) 548–559

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Activation of persulfate by CuOx@Co-LDH: A novel heterogeneous system for contaminant degradation with broad pH window and controlled leaching

MARK

Ali Jawadb, Jie Langb, Zhuwei Liaob, Aimal Khana, Jerosha Ifthikara, Zhanao Lva, Sijie Longb, ⁎ Zhulei Chenb, Zhuqi Chena, a Key Laboratory of Material Chemistry for Energy Conversion and Storage, Ministry of Education, Hubei Key Laboratory of Material Chemistry and Service Failure, School of Chemistry and Chemical Engineering, Huazhong University of Science and Technology, Wuhan 430074, PR China b School of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan 430074, PR China

G RA P H I C A L AB S T R A C T

A R T I C L E I N F O

A B S T R A C T

Keywords: PS activation Phenol degradation Broad pH window Diverse free radicals Scavenging effect

To meet the current challenges of waste water, an efficient, environment friendly and economical treatment process is always desirable for the degradation of toxic organic compounds. Herein, we reported a stable and active bimetallic CuOx@Co-LDH catalyst for the degradation of phenol using persulfate (PS) under mild conditions. The system of PS/CuOx@Co-LDH completely degraded 0.1 mM of phenol in less than 40 min while at the same time, all the other tested catalysts including catalyst precursors remained sluggish. In view of the practical application, CuOx@Co-LDH/PS system demonstrated consistent performance over a broad pH range (pH 5.0–12.0), during which minimum leaching could be detected. Moreover, good resistance to various inorganic anions was demonstrated in CuOx@Co-LDH/PS system. The consistent performance of CuOx@Co-LDH/PS system may originate from the constant buffer pH of catalyst (9.5 ± 0.8), minimum leaching and the generation of diverse free radicals (%SO4−, %OH or −%O2) under different pH conditions. Additionally, LDH structure offered strong synergetic interactions among active components which might also help in minimum leaching and good reactivity of CuOx@Co-LDH catalyst. The results of radical scavengers, EPR and XPS studies collectively proved

⁎

Corresponding author. E-mail address: [email protected] (Z. Chen).

http://dx.doi.org/10.1016/j.cej.2017.10.097 Received 25 May 2017; Received in revised form 16 October 2017; Accepted 17 October 2017 1385-8947/ © 2017 Elsevier B.V. All rights reserved.

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the mechanism of diverse free radicals, which was accounted to the redox cyclic reactions of active sites (Cu(III)–Cu(II)–Cu(III) and Co(III)–Co(II)–Co(III)) during degradation of phenol.

1. Introduction

[1,33,34]. These catalysts not only demonstrated excellent efficiency but also offered outstanding stability in bicarbonate activated hydrogen peroxide system (BAP). In this work, the prepared CuOx@Co-LDH catalyst shows better performance than their precursors of Co-LDH and Cu-LDH in PS activated system. Besides reactivity, the stability of CuOxCo-LDH catalyst in term of leaching was also remarkably improved. CuOx@Co–LDH catalyst was further studied for 1) consistent performance at a wide pH range (3.0–12), 2) generation of free radicals at pH-5, pH-7 and pH-10 and 3) influence of inorganic anions, particularly chlorides ions. Finally, free radical mechanism based on radical scavengers, EPR, ATRFTIR and XPS studies was proposed for the degradation of phenol and other target compounds.

The decrease of freshwater resources has motivated the researcher to recover water from contaminated sites. Among various technologies developed over years, advanced oxidation processes (AOPs) are considered more attractive due to the good mineralization efficiency and mild operating conditions [1–4]. In order to improve the working conditions of AOPs, modified technologies such as photo-Fenton’s, electro- Fenton’s, sono- Fenton’s and sono-photo Fenton’s have been established recently [5]. However, the strict pH requirement (∼3.0), energy input, difficulties in the storage and transportation of oxidant, metals release and post treatment of sludge make AOPs cost ineffective and non-ecofriendly [5]. In pursuit of alternative systems, peroxymonosulfate (PMS) or persulfate (PS) has got increasing research interests due to the generation of sulfate radicals (SO4%−) which possesses a higher oxidation potential (2.5–3.1 V) and a longer half-life (30–40 μS) than that of %OH radical (2.2–2.7 V, < 1 μS) [5–12]. Additionally, PS is cheaper than both H2O2 and PMS, and is more stable than H2O2 [13,14]. PS alone without activation presents very low oxidation potential toward organic contaminants. Different approaches such as heating, light, quinone and electron donors such as low valent metals (Fe0, Fe2+ & Ag+) were commonly used for PS activation [10,12,15–18]. However, the extra energy/chemicals input and the discharge of toxic metals into the environment are major limitations in these systems. Additionally, quinone is highly toxic and its introduction into contaminated water would not be recommended. On the other hand, mineral oxides (Fe2O3, α-FeOOH, MnO2, V2O3) as heterogeneous catalysts have got the advantage of minimum risk from soluble metal ions during PS activation [13,19]. However, the low efficiency, long contact time and high doses of PS (usually more than 100 times comparing to pollutants) are their major limitations for practical application. Co3O4 and supported cobalt catalysts (i.e. Co-MCM–41, Co-ZSM–5, Co-RMT, Co-SiO2, Co-SBA) appeared a good efficiency, but concerns about the leaching of carcinogenic cobalt are still not solved [20–25]. The search for alternative catalyst that can effectively activate PS without environmental risk remains a priority for the development of this technology. Recently, bimetallic oxides with spinal structures (CuFe2O4, CoFe2O4 and CuCo2O4) were identified as highly efficient catalysts than their single metals oxides in PMS system due to the synergetic coupling of active metals [26–28]. Considering the environmental and practical aspects, the incorporation of cobalt and copper onto a single support can provide a good choice for preparing efficient catalyst. Different kinds of supports (i.e. molecular sieves, mud, carbon, MgO or zeolites) were commonly used [22–24,29]. However, their crystalline structures facilitate catalyst leaching and poor stability due to the inhomogeneous distribution as well as limited interactions among active components. In this regard, the basic layered double hydroxide (LDH) compounds are very attractive because of the structural diversity (i.e. one can bring metals of interest in the desired ratio within the layered structure) [1,30]. More disperse composition, strong linkages among active sites, exchangeable anions, simple preparation methods and low costs are the additional features associated with LDH materials [1,30]. Additionally, LDH compounds have the ability to keep weak alkaline pH throughout the reaction which can benefit heterogeneous catalysts to reduce leaching. On the other side, weak alkaline pH favours the generation of diverse radicals (SO4%−, %OH and %O2−) which can help to minimize the influence of inorganic constituents in PS activated system [29,31,32]. In our previous work, we had applied these attractive properties of LDH in the fabrication of Co-supported LDHs catalysts

2. Materials and experimental section 2.1. Chemicals Phenol, sodium persulfate (Na2S2O8), tert-butyl alcohol (TBA), and methanol (MeOH) were obtained from Shanghai No.3 Reagent Factory (China). Other chemicals such as 1,4-benzoquinone (98.7%), cobalt nitrate hexahydrate (Co(NO3)2·6H2O, magnesium nitrate hexahydrate (Mg(NO3)2·6H2O), aluminium nitrate nanohydrate (Al(NO3)3·9H2O), copper nitrate trihydrate (Cu(NO3)2·3H2O) were of analytical grade and purchased from Sinopharm Chemical Reagent Co. Ltd (China). The EPR spin trap such as 5, 5-dimethyle-1-pyrolene N-oxide (DMPO) (98%) were purchased from Adamas Reagent Co. Ltd. 2.2. Preparation of powdered catalysts The Co-LDH uncalcined catalyst was prepared by a method as reported previously [1]. In general procedure, nitrate salts of Co2+, Mg2+ and Al3+ in certain ratio (Table S1) were drop-wisely added to a basic solution (2 M Na2CO3) with continuous stirring at 60 °C. The precipitate obtained after washing was dried overnight at 90 °C which was named as Co-LDH uncalcined catalyst. The Co-LDH uncalcined catalyst was heated at 500 °C for 5 h and was named as Co-LDH calcined. Similarly for comparative study, some control catalysts such as Cu-LDH, MgFeLDH and MgAl-LDH were also prepared by the above-mentioned method. For preparation of CuOx@Co-LDH catalyst, 0.5 g of Co-LDH calcined catalyst was dispersed in 1 mM of Cu(NO3)2·3H2O (15 mL) using sealed autoclave for 4 h at 100 °C. The resulting black suspension was first air dried in oven, calcined at 500 °C for 5 h and was named as CuOx@Co-LDH catalyst. For comparison, other catalysts such as FeOx@Co-LDH, NiOx@Co-LDH, MnOx@Co-LDH and CuOx@MgFeLDH were also prepared with the same procedure. 2.3. Experimental procedure 2.3.1. Contaminants degradation at the bench level Batch experiments for degradation of phenol, chlorophenols, nitro phenol and substituted benzene like compounds were conducted in glass bottles (30 mL) placed in a water bath with a magnetic stirrer at 30 °C. The reaction suspension was prepared by mixing PS (5.0 mM), catalyst (0.3 g/L) and 0.1 mM of target compounds in 20 mL deionized water. Periodically 2 mL sample was withdrawn with an interval of 10 min, filtered immediately through 0.2 μm filter and quenched with excess of ethanol prior to analysis for degradation. Experiments conducted at different initial pHs (not buffered) were performed at 549

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2.4. Characterization techniques

identical conditions. The initial pH was adjusted with 1 M NaOH or 1 M HCl. Anions or scavengers of appropriate amounts were added before the addition of PS.

Catalysts were characterized by different techniques such as X-ray diffraction analysis (XRD, X’Pert PRO), fourier transform infrared spectroscopy (FT-IR Bruker Equinox 55), Brunauer-Emmett-Teller (BET) analysis (Micromeritics ASAP 2020), atomic emission spectroscopy (MP-AES 4107, Agilent) and scanning electron microscopy (SEM Virion 200). Similarly X-ray photoelectron spectroscopy (XPS), ATRFTIR spectroscopy and Electron paramagnetic resonance (EPR) studies were conducted on V.G scientific ESCALAB mark II system, PerkinElmer (spectrum 100) and Bruker EMX 10/12 spectrometer respectively.

2.3.2. Analysis of contaminants The degradation of phenol and other target compounds was analyzed on a high performance liquid chromatography (HPLC FL–2200) equipped with a C18 (250 mm × 4.6 mm, 5 μm) column and a UV detector. A mixture of methanol and water (60/40, v/v) was used as the mobile phase at a constant flow rate (1 mL/min). The concentration of leached metal ions such as copper and cobalt was determined at the end of reaction by microwave plasma atomic emission spectroscopy (MP–AES4107, Agilent).

3. Results and discussion 3.1. Catalyst characterization

2.3.3. Identification of free radicals with EPR EPR studies were conducted using DMPO as the spin trapping agent. A solution containing 20 mM DMPO, 5.0 mM PS was prepared at pH 5.0, 7.0 and 10.0 ± 0.1, and then the catalyst was added to initiate the reaction. After 10 min of reaction, samples were taken and analyzed on a JEOL FA200 spectrometer at the room temperature. EPR parameters were adjusted at a radiation frequency of 9.147 GHz (X band), modulation frequency of 100 kHz, modulation width of 0.1 mT, sweep width of 20 mT, center field of 326.0 mT, scan time of 60 s, time constant of 0.03 s, and microwave power of 5 mW. ATR-FTIR study was conducted to characterize the interaction of PS and CuOx@Co-LDH as reported previously [26].

XRD patterns of CuOx@Co-LDH along with its precursors such as Co-LDH uncalcined, Co-LDH (calcined) and Cu-LDH (calcined) were shown in Fig. 1A. For Co-LDH (uncalcined), three peaks at 2θ ≈ 11°, 23°, 35° represented 001, 003 and 006 planes, while the doublet peaks at 2θ ≈ 61° and 62° indicated 110 and 113 planes of LDH compounds, respectively [33–35]. After calcination, the LDH structure was destroyed and MgO phase associated with two broad peaks at 2θ ≈ 43° and 63° (JCPDS-079-0612) was appeared. XRD spectra of CuOx@CoLDH shows peaks for three major phases: 1) 2θ ≈ 11°, 23°, 35.5°, 38.8° for LDH, 2θ ≈ 43°, 63° for MgO and 2θ ≈ 32°, 48° for CuO [30,33,36,37]. However, this separate phase of CuO cannot be found in Fig. 1. (A) XRD pattern of CuOx@Co-LDH and their precursors. (+CuO phase) (B) SEM image of Co-LDH and (C) SEM image of CuOx@Co-LDH catalyst showing morphological changes after Cu impregnation.

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redox metals as evident from the common band at 465 cm−1 in all studied materials. In order to find the oxidation states of cobalt and copper, X-ray photoelectron spectra (XPS) were conducted for both fresh and used CuOx@Co-LDH catalyst (Fig. 2). The survey scan of CuOx@Co-LDH revealed the presence of Mg 1s, Cu 2p, Co 2p, O 1s, C 1s and Al 2p (Fig. 2A). In Fig. 2B, the Cu 2p3/2 core region in fresh CuOx@Co-LDH catalyst appeared at the binding energy of 933.60 eV along with intense shakeup peaks. In literatures, metallic copper or Cu2O appears below 933 eV with no satellite peak, while CuO usually appears at binding energies higher than 933 eV (933.5–935.1 eV) along with shakeup peaks [39–41]. This information strongly recommended Cu2+ as CuO phase in fresh CuOx@Co-LDH catalyst, which was consistent with XRD findings. In used CuOx@Co-LDH catalyst, some of the Cu 2p3/2 peak region shifted to a higher binding energy (935.06 eV) which suggested that some of Cu2+ ions may appeared more electron deficient that could facilitate the generation of Cu3+ during the reaction [42]. Regarding Co (inset of Fig. 2C and D), no signal was detected in both fresh and used CuOx@Co-LDH catalyst. Probably the impregnation of CuOx encircled the very limited Co species (0.6% in weight) and thus made it deep and beyond the detection limit of XPS (upper 5 nm) [28]. As CoLDH was used as the support during the preparation of CuOx@Co-LDH catalyst, Co-LDH was used for XPS analysis to confirm the oxidation state of cobalt (Fig. 2C). The Co 2p3/2 and 2p1/2 levels appeared at 780.9 and 795.9 eV, respectively, with a satellite peak at 786.4 eV, which confirmed the Co2+ and Co3+ states of cobalt in Co-LDH calcined catalyst [28]. Accordingly, this indirect evidence suggested that Co may also exist in the states of Co2+ and Co3+ in CuOx@Co-LDH catalyst. Together with the elemental analysis, basic properties of powdered catalysts were summarized and shown in Table S1.

Cu-LDH. Thus we believe the Cu components in CuOx@Co-LDH catalyst existed in the form of both highly dispersed one in LDH structure and a separate phase of CuO. CuOx@Co-LDH catalyst re-gained some of the LDH structure as evidenced from weak 001, 003 and 006 reflections due to the memory effect after the impregnation process [30]. However, their memory effect was weak by comparing with Co-LDH calcined catalyst, which re-gained all principal peaks of LDH after suspending in aqueous medium for 1 h (Fig. S1). Regarding cobalt, no separate peak of cobalt in any form was found in both Co-LDH calcined catalyst and CuOx@Co-LDH catalyst due to the very small loading amount (0.6%). Also, LDH compound generally produces a highly dispersed mixture of different components after calcination, and this property may be responsible for not detecting any cobalt peak [37]. On the other side, SEM image (Fig. 1B) confirmed that Co-LDH particles appeared in the typical hexagonal LDH symmetry. After hydrothermal impregnation of Cu as CuOx@Co-LDH catalyst, the CuO phase encircled all of the Co-LDH particles, making it extremely rough to be clearly visualized in Fig. 1C. The surface area of Co-LDH (92 m2/g) and Cu-LDH (84 m2/g) was very similar to that of MgAl-LDH (87 m2/g) (Table S1). On the other hand, CuOx@Co-LDH catalyst showed a significantly higher surface area (131 m2/g) due to hydrothermal treatment. FTIR analysis was further applied to characterize CuOx@Co-LDH as well as its precursors (Fig. S2). In the FTIR spectrum, shifting of certain absorption bands from MgAl-LDH taken as reference standard was found in the wavelength region below 1500 cm−1 (Fig. S1). For example, the CO32− band at 1379 cm−1 in MgAl-LDH was shifted to a higher wavelength (1386 cm−1) after introducing Cu in CuOx@CoLDH. The bands at 664 and 420 cm−1 in MgAl-LDH were assigned to MOH and O-M-O vibrations respectively [38]. After introducing redox metals in MgAl LDH structure, the O-M-O vibrations shifted to higher wavelength. For instance, O-M-O vibrations first shifted to 450 cm−1 both in Co-LDH and Cu-LDH and later after impregnation of Cu as CuOx@Co-LDH, these vibrations were further shifted to 483 cm−1. Interestingly, M-OH vibration were not affected after introducing these

3.2. Catalytic activity of different LDH based catalysts Initially different LDH based catalysts (i.e. CuOx@Co-LDH, Fig. 2. XPS analysis of CuOx@Co-LDH showing the electronic state of Cu and Co (A) survey scan of elements in CuOx@Co-LDH catalyst (B) binding energies of Cu in CuOx@Co-LDH catalyst before the reaction (C) binding energies of Co in Co-LDH catalyst (inset showing the binding energies of Co in CuOx@Co-LDH catalyst after the reaction.

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dose for Co-LDH or Cu-LDH catalysts (0.4 g/L). Considering the Co-LDH or Cu-LDH catalysts (0.4 g/L) possessed larger numbers of active sites than CuOx@Co-LDH (0.2 g/L), the huge difference of their reactive rates should be attributed to the synergetic effect of Co and Cu in CuOx@Co-LDH, while the difference of surface area may not be the reason for enhanced activity of CuOx@Co-LDH. Additionally, this fact can be recognized from the much lower degradation of physical mixture of Cu-LDH, Co-LDH and CuO (2 mg each Fig. 3B). In order to evaluate the stability of these catalysts, the leaching of active components was also determined (Table S1). In contrast to Co-LDH or Cu-LDH, CuOx@Co-LDH catalyst appeared more stable as no leaching of Co2+ and 0.08 ppm leaching of Cu2+ ions were detected at the end of reaction. To check the possible activation of PS with metals ions in homogeneous phase, 0.32 ppm Cu2+ plus 0.20 ppm Co2+ ions were used (Fig. 3B, or even Cu2+ ions of 64 ppm in Fig. S4). Results of these experiments highlighted the heterogeneous activation of PS as almost no phenol was degraded using 0.32 ppm Cu2+ plus 0.20 ppm Co2+. Considering these promising properties (excellent efficiency and stability), CuOx@Co–LDH catalyst was further studied to explore its practical application in PS activated system.

FeOx@Co-LDH, FeOx@Cu-LDH, NiOx@Co-LDH and MnOx@Co-LDH) with combination of different redox metals were tested in PS activated system for degradation of phenol as target compounds (Fig. 3A). Among them, CuOx@Co-LDH catalyst demonstrated an excellent efficiency with nearly 100% phenol removal in less than 40 min. Other catalysts such as FeOx@Co-LDH, FeOx@Cu-LDH, NiOx@Co-LDH and MnOx@Co-LDH were comparably sluggish in activating PS, and offered only 10–30% phenol removals under identical conditions. Meanwhile the CuOx@Co-LDH precursors like MgAl-LDH, Cu-LDH, Co-LDH (Fig. 3A and B) or the physical mixture of Cu-LDH + Co-LDH catalysts offered only 7%, 34%, 37%, and 25% phenol removal, respectively. Additionally, CuOx@Co-LDH or PS alone and Cu2+, Co2+ ions as homogeneous phase catalyst (0.32 + 0.20 mg/L respectively) (Fig. 3B) showed negligible degradation (≤10%). Further, CuOx@Co-LDH catalyst was found even more reactive than the commonly reported catalysts in term of degradation rate constant during activation of PS (Table 1) [14,19,39,43–47]. The higher degradation rate constant and excellent activity of CuOx@Co-LDH catalyst may be possibly attributed to; (1) the increase of surface area (Table S1) which would not only facilitate better adsorption but also expose more active sites to activate PS and to generate more radicals for catalytic reactions; (2) improved stability in CuOx@Co-LDH catalyst (less leaching, Table S1); (3) high redox activity from the synergetic coupling of mixed valence transition metals. Firstly, the BET area as summarized in Table S1 indicated that surface area might not be the dominant parameter effecting the catalytic activities of these catalysts. The surface area of CuOx@Co-LDH, Co-LDH and Cu-LDH/PS were 131, 92 and 84 m2/g (MgAl-LDH alone: 87 m2/g). However, the apparent rate constant with CuOx@Co-LDH/PS was 19.77 and 17.45 times higher than Co-LDH/PS and Cu-LDH/PS respectively. Further, Fig. 3C (original data in Fig. S3) showed the relationship between apparent rate constants and catalyst doses for the three catalytic systems. CuOx@Co-LDH catalyst at a smaller dose (0.2 g/L) achieved a much higher reaction rate than the case of a larger

3.3. Factors influencing the degradation of phenol 3.3.1. Effect of CuOx@Co-LDH dosage Effect of different amounts of CuOx@Co-LDH catalyst (0, 0.2, 0.25, 0.30, 0.35 & 0.40 g/L) were evaluated on degradation of phenol at a fixed amount of PS (5 mM) and phenol (0.1 mM) (Fig. 4A). These results indicated that PS without catalyst could only degrade less than 5% of phenol while adding 0.2 g/L of catalyst, the degradation of phenol abruptly increased to 78%. By increasing the amount of CuOx@Co-LDH catalyst, the degradation of phenol continuously increased and reached 100% at 0.3 g/L catalyst. The active sites for PS activation increased by increasing the amount of catalyst. Such condition are favorable for PS Fig. 3. (A) Degradation of phenol with different LDH based catalysts, (B) Activity of CuOx@Co-LDH catalyst during degradation of phenol (C) Variation of pseudo first order degradation rate constant of phenol with catalysts doses.

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Table 1 Degradation rate constants of organic compounds during activation of PS with different catalysts. Catalyst

Conditions

Target compound

Maximum removal/Reaction time

Degradation rate constant (min−1)

Ref.

CuOx@Co-LDH

Phenol 0.1 mM 0.1 mM

100%/0.30 h

0.170

This work

80%/2 h

–

[39]

Magnetic nanoparticle (MNP)

0.3 g/L catalyst, PS 5 mM, 30 °C 0.3 g/L catalyst, PS 5 mM, 30 °C MNP 1 g/L, PS 2 mM, 25 °C

90%/4 h

0.010

[43]

Carbon nanotubes (CNT)

CNT 0.2 g/L, PS 6.2 mM, 25 °C

100%/2 h

0.044

[44]

Annealed Nano diamond (AND)

AND 0.2 g/L, PS 6.5 mM

100%/0.45 h

0.084

[14]

CuBi2O4

0.1 g/L catalyst, PS 0.37 mM

100%/5 h

0.015

[45]

1g/L catalyst, PS 12.4 mM, 25 °C

PCB 2.5 μM Phenol 0.213 mM Phenol 0.21 mM H-benzotriazol (BTZ) 0.02 mM DCP 0.02 mM

34, 5.7 and 38%/3 h

0.0013, 0.0018, 0.007

[46]

0.624 g/L catalyst, PS 10 mM, 30 °C 0.05 g/L catalyst, PS 2 mM, 25 °C

DCP 0. 6.1 mM PCB 28 3.9 μM

96.7%/3 h

0.022

[47]

82%//4 h

0.0072

[19]

CuO-Fe3O4

Fe

2+

Fe2O3 nZVI CuO-Fe3O4 V2O3

activation to generate free radicals (%OH & SO4%−) and thus increased the degradation efficiency of catalyst [48]. By further increasing the amount of the catalyst, the degradation efficiencies did not changed, which was different from some studies suggesting the adverse effect owing to the competitive consumption of radicals with excess amount of catalysts and diffusion limitation phenomenon [26,49].

of phenol was decreased to 20 min by increasing the amount of PS from 4 mM to 8 mM. The generation of more free radicals as also reported at a high dose of PS may contribute for the faster degradation of phenol [50]. 3.3.3. Effect of phenol concentration The degradation of different initial phenol amount (0.05, 0.1, 0.15, 0.20 & 0.25 mM) were also evaluated at optimized amount of PS (5 mM) and CuOx@Co-LDH catalyst (0.3 g/L). As shown (Fig. 4C), small amounts of phenol (5–10 ppm) were completely degraded in 20–40 min respectively. However, by increasing the initial amount of phenol (15–25 ppm), the degradation efficiency of CuOx@Co-LDH catalyst dropped continuously to 80% at 15 ppm and then to 60% at 25 ppm. When excess amounts of phenol reacted with fixed number of free radicals/active sites, the increasing competition between target

3.3.2. Effect of PS dosage The performance of CuOx@Co-LDH catalyst was also evaluated at different amounts of PS (0, 2, 4, 5, 6 & 8 mM) at a fixed amount of phenol (0.1 mM) and catalyst (0.3 g/L) (Fig. 4B). These results indicated that CuOx@Co-LDH catalyst alone has no reactivity and the addition of small amount of PS (2 mM) abruptly increased the degradation of phenol to 80% in 40 min. By further increasing the amount of PS, degradation was accelerated. For instance, the degradation time

Fig. 4. influence of (A) CuOx@Co–LDH dosage (B) PS amount (C) phenol concentration (D) temperature during degradation of phenol.

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coordination sites for complexation with oxidants which in turn accelerate the redox properties of heterogeneous catalysts during catalytic reactions [26,52]. Here, ATR-FTIR was applied to confirm any possible complexation between PS and CuOx@Co-LDH catalyst (Fig. S5A and B). IR bands in the range of 900–1340 cm−1 came from the stretching vibration of SeO bond of either HSO5− or SO42− [53]. A clear shift of about 10 cm−1 was noticed for the bands of PS at 1299, 1294 and 1060 cm−1 when catalyst was added. This shift was ascribed to the complexation between the active sites of CuOx@Co-LDH and PS. In addition, when water spectrum was subtracted as the background, the broad IR band of CuOx@Co-LDH catalyst at 3633–3178 cm−1 indicated the presence of considerable amount of surface –OH groups (Fig. S5B). However, when PS was added, these peaks became narrow and their intensities were also decreased. These results thus indicated that some of –OH groups were replaced when PS was bonded with the active sites of CuOx@Co-LDH catalyst. Similarly, the inner-sphere complexation was further evidenced from the negligible influence of increasing ionic strength during degradation of phenol (Fig. S6). Because changes in ionic strength (changes in electrostatic interactions) had no effect on the inner-sphere complexation between PS and active sites of CuOx@Co-LDH catalyst, it is clear now that the dominant interaction between PS and active sites of CuOx@Co-LDH catalyst was through covalent bonds rather than through electrostatic attractions. Due to this reason, CuOx@Co-LDH/PS system offered constant performance over a wide pH range, regardless of pHZPC (Fig. 5D) which can be used to measure the strength of electrostatic interactions at any studied pH. Moreover, the uncharged oxygen atom of PS (SO3-O-O-SO32−) is expected to be coordinated with the active sites of CuOx@Co-LDH catalyst to produce powerful free radicals. Further, changes in pH as well as leaching of active components were also monitored (Fig. 5B and C) during the reactions. As shown in Fig. 5B, the degradation of phenol was interestingly correlated to the final pH of solution. For example, the eventual pH of solution remained stable at 9.5 ± 0.8 for all initial pHs (5.0–10.0). Thus the constant performance in the initial pHs (5.0–10.0) may be partly attributed to

compounds and free radicals/active sites negatively influenced the degradation efficiency of CuOx@Co-LDH. 3.3.4. Influence of reaction temperature The effect of temperature (25, 30, 35, 40 °C) on CuOx@Co-LDH/PS system was listed in Fig. 4D. The degradation of phenol continuously increased with increasing temperatures. For instance, the degradation of phenol improved from 69% to 100% after 20 min of reaction time when the temperature increased from 25 °C to 40 °C. Thus heating was beneficial for activating PS to achieve quick degradation of phenol. From the apparent rate constants at different temperatures, the calculated activation energy was 52.90 kJ/mol, consistent with the previously reported PS based system for degradation of phenol [22]. 3.3.5. Performances of CuOx@Co-LDH/PS system at different pHs It is well known that the performances of PS or PMS based technologies are mostly dependent at solution initial pHs [39]. Here we also investigated the influence of initial solution pH over a wide range (pH 3.0–12.0, not buffered) using CuOx@Co-LDH/PS system during phenol degradation. As shown in Fig. 5A, the degradation of phenol abruptly increased from 40% at pH 3.0 to 88% at pH 5.0 in 40 min of reaction time. After this abrupt change, the CuOx@Co-LDH/PS system efficiency became nearly constant at ≥98% from pH = 5.0 to 12.0. It is worth to mention that, these results are different from the previously reported studies. For example, during heat–assisted PS degradation of methyl tert–butyl ether (MTBE) in the range of pH (2.5–9.0), the highest degradation was found at pH 2.5 [51]. Similarly, Zhang et al. [50], observed highest degradation at pH 3.0 for PS/CuFe2O4 system. Three factors: 1) interactions of catalysts, contaminants or oxidants, 2) changes of solution pHs with time and 3) leaching of active components were considered to explain the broad pH range activity of CuOx@Co-LDH/PS system. Firstly, change in solution pH can effect the interactions between catalyst particles, contaminants and oxidant molecules due to the protonation and deprotonation of surface hydroxyl groups. Hydroxyl groups at the surface of catalysts provide

Fig. 5. (A) Influence of pH on the efficiency of CuOx@CoLDH/PS system during degradation of phenol (B) Variation of pH during reaction (C) Cu2+ leaching at different pH reactions in the CuOx@Co-LDH/PS system (D) pHZPC measurement of CuOx@Co-LDH by zeta potential method.

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(1.6–7.7 × 107 M−1s−1) [54]. Similarly, tert-butanol (TBA) reacts with hydroxyl radical about 1000 times faster (3.8–7.6 × 108 M−1s−1) than %SO4− radical (4.4–9.1 × 105 M−1s−1) and thus these scavengers can be used to differentiate between %SO4− and %OH radicals [54]. In order to get sufficient scavenging, the maximum required amount of these scavengers were first optimized at pH-10.0 (Fig. S7). As demonstrated, the 2.5 M ethanol and 0.5 M TBA already achieved the maximum inhibition effect. At pH 5.0 as shown in Fig. 6A and Table 2, 90% of phenol was degraded in 40 min when no scavenger was added. By adding TBA, the degradation of phenol dropped to 71% owing to the scavenging of %OH radicals. The degradation of phenol decreased significantly to 30–33% in the presence of ethanol or mixture of ethanol and TBA, which was accounted for trapping of both %SO4− and %OH radicals (Table 2). Further, 1,4–benzoquinone (BQ) alone, a popular scavenger for super oxide (%O2−) radicals [1], inhibited 28% degradation alone. On the other hand, the scavenging of ethanol (57%) or ethanol + TBA (60%) was nearly equal to ethanol + TBA + BQ (63%) at pH 5.0 (Fig. 6A and Table 2), indicating superoxide could be neglect in this case. The findings thus suggested the collective contribution of %SO4− and %OH radicals in the degradation of phenol at pH 5.0. At pH 7.0, the degradation of phenol decreased further with TBA from 100% to 69% (Fig. 6B). Similarly, the degradation of phenol further dropped from 100% to 35–37% in the presence of ethanol alone or mixture of ethanol and TBA under identical reaction conditions. In the presence of TBA at pH 10.0, the degradation of phenol dropped to 54% while ethanol alone or mixture of ethanol and TBA accounted for 71–73% decrease in the degradation of phenol (Fig. 6C). This study thus indicated that the scavenging role of TBA accounted for %OH radicals gradually increased from 19% at pH 5.0 to 46% at pH 10.0, which was accounted for the transformation of %SO4− radicals into %OH radicals (Eqs. (1), (2)) [55]. However, TBA alone did not get complete scavenging to become similar like ethanol at any studied pH which suggested the contribution of both %SO4− and %OH radicals in the degradation of phenol under all pHs (Table 2). The generation of %O2− radicals was well established in alkaline pHs (Eqs. (3) and (4)) through the reaction of HO2− and S2O82− [56,57]. Further, Fang et al. [43] suggested %O2− radicals as driving agent for activation of PS. The experiments

the buffer capacity of CuOx@Co-LDH catalyst [39]. However, this buffer ability was negligible when the initial pH was 3.0 or above 10.0. On the other hand, leaching of metals ions is commonly accounted for the loss of activity in heterogeneous catalysts. The lower activity of CuOx@Co-LDH catalyst at pH 3.0 may be due to the leaching (4.32 ppm for Cu2+ and 0.16 ppm for Co2+) in strong acidic medium (Fig. 5C). XRD characterization for catalyst that reacted at pH-3 and pH-7 gives additional strong evidences (Fig. S1). The typical peaks for MgO phase (2θ ≈ 43°, 63°) and LDH phase (2θ ≈ 11°, 23°, 35.5° & 38.8°) were completely disappeared while CuO phase (2θ ≈ 42°, 48°) got weaken. Further, the adverse effect of H+ ions may also contribute at lower pHs which can decrease the chances of complex formation on the catalyst surface [15]. PS activation in highly alkaline pHs (11.0–12.0) is a common phenomenon [39]. We got less than 15% phenol removal by conducting control experiments in highly alkaline pHs (11.0–12.0) which suggested small contribution of base activation in this pH range. The leached Cu2+ ions from pH 5.0 to 12.0 was gradually decreased from 1.17 to 0.08 ppm (Fig. 5C). By contrast, no leaching of Co2+ ions were detected throughout the studied pH range. Thus we concluded that the consistent performance of CuOx@Co-LDH catalyst over a wide range of pH (5.0–12.0) is collectively related to the inner-sphere complexation of oxidant and catalyst, buffer pH and stability of catalyst. 3.4. Generation of diverse free radicals at different pHs Previous studies indicated that %SO4− and %OH radicals could be generated in PS/PMS system (Eqs. (1) and (2)) [26,39,43,52,53]. % SO4− + H2O→SO42−+% OH+H+

(1)

% SO4− + OH−→SO42−+% OH

(2)

Here, we performed free radical quenching studies to identify the contribution of different free radicals in CuOx@Co-LDH/PS system at different pHs. Ethanol and tertiary butyl alcohol (TBA) were used as selective scavengers owing to the significant difference in their rate constants with %SO4− and %OH radicals. For example, the second rate order constant of ethanol toward %OH radical (1.2–2.8 × 109 M−1s−1) is approximately 50-fold faster than that of %SO4− radical

Fig. 6. Effect of scavengers on the activity of CuOx@CoLDH/PS system during phenol degradation. (A) Effect of radicals scavengers at solution pH = 5.0 (B) Effect of radicals scavengers at pH = 7.0 (C) Effect of radicals scavengers at pH = 10.0

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Table 2 Summary of the phenol removal efficiency by scavengers. pH

5.0 7.0 10.0

% removal without scavenger

90 100 100

% trapping with single scavenger

% trapping with multiple scavengers

Dominant radical

TBA

Ethanol

BQ

Ethanol + TBA

Ethanol + TBA + BQ

19 31 46

57 65 71

28 55 64

60 67 73

63 76 86

SO4%−,%OH SO4%−, O2%−, %OH SO4%−, O2%−,%OH

Conditions: [phenol] = 0.1 mM, [PS] = 5 mM, amount of CuOx@Co-LDH = 0.3 g/L, [TBA] = 0.5 M, [MeOH] = 2.5 M, BQ = 2 mM, T = 30 °C, Time = 40 min.

As shown in Fig. 8A, CuOx@Co-LDH/PS system offered good resistance to the presence Cl− ions. The degradation of phenol remained constant at different concentration of Cl− ions (0–200 mM). The impact of Cl− ions in SO4%− radicals based systems is controversial among researchers [26,49–51]. For examples Zhang et al. [26] and Jing et al. [57], reported the negative impact of Cl− ions even at a very small concentration. On the other side, Gu et al. [58], found a greater scavenging effect at higher concentrations and [59], observed positive influence on degradation. Generally, both SO4%− or %OH radicals tends to be rapidly scavenged by halogens ions (Eqs. (5) and (6)).

conducted with the addition of BQ demonstrated significant scavenging of phenol degradation at pH 7.0 and 10.0 (Fig. 6B and C & Table 2). However, based on the non-selectivity of BQ toward %O2− and low reactivity toward phenol, this scavenging may not be totally accounted for superoxide radicals. Further, experiments conducted with the mixture of scavenger (BQ + ethanol + TBA) gave the clear contribution of superoxide in the degradation of phenol. The scavenging of ethanol + TBA + BQ (76%) was 9% higher than ethanol + TBA (67%) at pH 7.0, and this difference arrived 13% at pH 10.0 (73% for ethanol + TBA and 86% for ethanol + TBA + BQ). Thus we concluded the major role of %SO4− and %OH, while %O2− was also accounted as reactive species for degradation of phenol at pH 7.0 and 10.0 (Table 2).

SO4%− + Cl− ⇌SO42− + %Cl = (2.1 ± 0.1) ×108 M−1s−1

Kf = (3.2 ± 0.2) ×108 M−1s−1,

Kr (5)

S2 O82 − + 2H2 O→ HO−2 + 2SO24− + 3H+

(3)

Cl− + %HO ⇌%HOCl− M−1s−1

HO2− + S2O82−→% SO4− + SO42− + % O2− + H+

(4)

As a result of scavenging reactions the generated new free radicals such as %Cl or %HOCl− offered negative impact on degradation because of its smaller oxidative potential [58,60,61]. In this context, the consistent performance of CuOx@Co-LDH/PS system might be originated from superoxide because of its low potential to react with increasing concentration of Cl− ions (Eq. (7)) [61].

To further confirm the identity of these ROS at different pHs, EPR spectroscopy was conducted using DMPO as the spin trapping reagent. As shown in Fig. 7A, the hyperfine splitting constant for DMPO-OH (aH = aN = 14.9 G) and DMPO-SO4 (aN = 13.2 G, aH = 9.6 G, aH = 1.48 G, aH = 0.78 G) were consistent with the previously reported data for %OH and %SO4− [10,27]. In the case of DMPO–OH, a clear quartet peak appeared under all conditions (pH 5.0, 7.0 and 10.0) for %OH radicals. However, its intensity apparently increased with the increase of pH, which indicated the transfer of %SO4− radicals into %OH radicals. The short half–life of DMPO–OOH/−O2 adduct (< 1 min), lower reactivity of superoxide toward DMPO with (k = 10−18 M−1S−1) and the presence of %SO4− and %OH radicals may cause problems to distinguish %O2− radicals in CuOx@Co-LDH/PS system using EPR [33].

Cl− + %O2

−

Kf = 4.2 ×109 M−1s−1,

Kr = 6.10 ×109 (6)

→ < 0.014 M−1s−1

(7) −

−

NO3 , CH3COO and HCO3− Others anions such as (Fig. 8B) also offered marginal scavenging effect in the first 30 min. However, their scavenging effect became less important on extending the reaction time for 1 h. For example, HCO3− can react strongly with both %SO4− and %OH radicals (k > 106 M−1s−1) (Eqs. (8) and (9), and thus caused some inhibitory effect in the initial 30 min [62]. However, on extending the reaction time (1 h), the degradation occurred smoothly until all of the phenol was degraded. Similar trend was observed for all other anions accept Br− and ClO4− which offered no inhibitory effect like Cl− ions. Again the generation of %O2- in CuOx@Co-LDH/PS system probably played an important role owing to the smaller scavenging potential toward inorganic anions (Eqs. (10) and (11)) and thus kept the degradation smoothly [63]. SO42−,

3.5. Influence of inorganic anions Inorganic anions such as Cl−, Br−, NO3−, CO32−, HCO3−, CH3COO− etc. are popularly known scavengers of free radicals. The natural occurrence of these anions in most of the treated water demonstrates tremendous impact on the efficiency of AOPs. As these anions offered different scavenging potentials for different free radicals. Therefore, the presence of various free radicals (%SO4−, %OH and %O2−) in CuOx@Co-LDH/PS system would be beneficial to offer minimum scavenging for different anions.

%OH + HCO3− →%CO32

−

+ H2O

K = 8.5 ×106 M−1s−1

%SO4− + HCO3− →SO42− + %HCO3− 8.4

(8)

K = 1.6 ×106 M−1s−1 at pH (9)

Fig. 7. (A) EPR spectra in CuOx@CoLDH/PS system at different pH with DMPO (B) Generation of SO4%− and %OH radicals in different catalyst/PS system.

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Fig. 8. Effect of inorganic anions on the efficiency of CuOx@CoLDH/PS system (A) Cl− ions in the range of 5–100 mM (B) 5 mM of different anions.

HCO3− + %O2− →HO2− + CO32−

to a much lower reduction potential of Cu2+/Cu+ (0.15 V) [39]. To verify this, we performed a control experiment (Fig. S4) where Cu+ resulted from the reduction of Cu2+ by ascorbic acid appeared with a better efficiency than Cu2+ + PS system, and thus confirmed its lower activation ability for PS [39]. Further, some previous works also confirmed that Cu2+/PS system needed a longer time for the degradation of organic compounds [64]. Secondly, XPS results (vide infra) indicated no apparent changes in the valence of copper before and after reaction (Fig. 2B and D). Therefore, much published work and our own results indicated that activation mechanism based on Cu3+/Cu2+ redox pair would be more suitable [26,39] because such reactions are thermodynamically feasible to activate PS (Cu3+/Cu2+, 2.3 V) [39]. Regarding Co (inset of Fig. 2C and D), although we did not detected it both in fresh and used CuOx@Co-LDH catalyst, the indirect evidence from XPS analysis conducted for Co-LDH (Fig. 2C) revealed Co2+ and Co3+ states. Additionally, ATR-FTIR analysis discussed in detail (vide infra) confirmed the generation of complex between the active sites of CuOx@Co-LDH (^Cu2+/^Co2+) and PS (SO3-O-O-SO32−) during the activation process (Eq. (12)). The activated complex on dissociation produce %SO4− radicals and high valence state of active sites (Eqs. (13) and (14)). The high valence states (^Cu3+ or ^Co3+) in the absence of strong chelating groups, react directly with water/HSO5− and thus regenerate the active sites along with %OH or %SO5− radicals (Eqs. (15)–(18)) [65].

K = 0.04 M−1s−1 at pH 10.1 (10)

%O2− + CH3COO− → < 0.06 M−1s−1 at pH 10

(11)

For other tested compounds (Fig. S8), the degradation profile of CuOx@Co-LDH/PS system followed the order of phenol > 2,4,6-TCP > 2-CP ∼ 2,3-DCP > 4-nitrochlorobenzene > nitrobenzene∼ pnitrophenol and thus indicate the selectivity trends in degradation towards phenol or chlorophenols (2-CP,2,3-DCP,2,4,6-TCP) rather than substituted benzene like compounds. The degradation of 2-CP and 2,3DCP was slightly lower than that of phenol under similar conditions of pH 7.0. This behavior may be related to their smaller pka values (8.56 and 7.60 respectively) than phenol (10.0). Such differences in pka values would influence the number of deprotonated 2-CP, 2,3-DCP or phenol, which may possibly involve in interactions with catalyst during degradation. Fig. S9 showed the HPLC chromatograms of phenol degradation after different time interval in CuOx@Co-LDH/PS system. In the start of reaction, phenol and PS showed their characteristic peaks. However, the peak of phenol decreased with time and a new peak identified as benzoquinone (BQ) was emerged. The BQ peak first increased in intensity (20 min) and then start decreasing without producing additional peaks until phenol and BQ peaks were disappeared completely in 30 min. This data confirmed that our degradation mechanism is not very different from those published works that the oxidation of phenol was through the oxidation of hydroxyl group, leading to the product of benzoquinone, and then mineralized to other products.

≡ Cu2 +/ ≡Co2 + + SO3−O−O−SO32 − → ≡Cu2 +/ ≡Co2 +⋯SO3−O−O−SO32 − (12) ≡Cu2+ ⋯SO3-O-O-SO32− →≡Cu3+ + %SO4− + SO42

3.6. Possible PS activation mechanism

→≡Co

3+

+ H2O →≡Cu

2+

+ %OH + H

3+

+ H2O →≡Co

3+

+ %OH + H

≡Co

Here, the radical scavengers and EPR studies first confirmed the presence free radical mechanism. Generally free radical mechanism is based on the transfer of electron from transition metals toward PS/PMS during activation process. For example, Li et al. [48] and Zhang et al. [50] proposed the generation of %SO4− radicals from oxidation of Cu+ and Fe2+ which were supposed the pre-reduced states of Cu2+ and Fe3+ in PS activated system. In contrast, Zhang et al. [26] and Lei et al. [39] proposed the oxidation of ^Cu2+ to ^Cu3+ in PMS/PS system after producing complex at the surface of CuFe2O4 and CuO-Fe2O4 catalysts. These studies thus suggested two major mechanisms (i.e. Cu2+/Cu+ v.s. Cu3+/Cu2+) in PMS/PS activation using copper based catalysts. In contrast, the activation of PS/PMS with cobalt is simple, associated with single redox cyclic Co3+/Co2+ reactions [27–29,57,60]. For example, Anipsitakis and Dionysiou [54] accounted the high activity of cobalt ions in PMS system for its high reduction potential (Co3+/Co2+ 1.92 V) and thermodynamic feasibility to regenerate its active Co2+ state. The activation mechanism based on the redox cycle of Cu2+/Cu+ might not be dominant in CuOx@Co-LDH/PS system due to following facts. Firstly, such reactions are thermodynamically not feasible owing

%SO4−

⋯SO3-O-O-SO32−

2+

≡Cu ≡Co

≡Cu ≡Co

3+

3+

+

+

+

→≡Cu

HSO5−

→≡Cu

2+

2+

+

%SO5−

+ %SO5

−

(13) (14)

(16) +

(17)

+

(18)

+H +H

+ SO4

2−

(15)

+

−

+ HSO5

3+

−

The significant lower activity of Co-LDH and Cu-LDH catalysts (Fig. 3B) than CuOx@Co-LDH demonstrated that possibly the presence of both these metals encourage the transfer of electron during complexation with PS. As a result more %SO4−/%OH radicals were produced as also evident from the much stronger EPR signals associated with CuOx@Co-LDH/PS system than Co-LDH/PS or Cu-LDH/PS system (Fig. 7B) [53]. Among generated radicals, %SO5− (Eqs. (17) and (18)) and %O2− (Eqs. (3) and (4)) has much weaker oxidizing ability toward organic pollutants. Additionally, the information collected from the radical scavenger’s studies suggested that super oxide has small role in the degradation of phenol. Therefore, it is thus concluded that both % SO4− and %OH radicals would be the dominant radicals in degradation of phenol (Eqs. (19) and (20)). 557

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%SO4− + phenol →degradation products

(19)

%OH + phenol →degradation products

(20)

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4. Conclusion In the present study we reported a simple, highly efficient and environment friendly CuOx@Co-LDH/PS system. CuOx@Co-LDH/PS offered an excellent efficiency for degradation as well as remarkable stability over a wide pH range, and also presented minimum scavenging for various inorganic anions. The main factors responsible for the consistent performance of CuOx@Co-LDH/PS system were (1) basic properties of LDH catalyst which offered natural basic buffer pH (9.50) irrespective of initial pH in the range of pH 5.0–10.0 (2) minimum catalyst leaching and (3) the generation of various active species (% SO4−, %OH and %O2−) under different initial pHs (5.0, 7.0 and 10.0). Finally based on XPS findings, ATR-FTIR and radical confirmation techniques, a suitable free radicals mechanism through redox cycle of Cu(III)-Cu(II)-Cu(III) and Co(III)-Co(II)-Co(III) was proposed for activation of PS. Acknowledgements This work is financially supported by the National Science Foundation of China (Grant No 21671072). The authors also thank the Analytical and Testing Center of Huazhong University of Science and Technology for helping in characterization of material. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.cej.2017.10.097. References [1] A. Jawad, Y. Li, X. Lu, Z. Chen, W. Liu, G. Yin, Controlled leaching with prolonged activity for Co–LDH supported catalyst during treatment of organic dyes using bicarbonate activation of hydrogen peroxide, J. Hazard. Mater. 289 (2015) 165–173. [2] L. Zhu, Z. Ai, W. Ho, L. Zhang, Core–shell Fe–Fe2O3 nanostructures as effective persulfate activator for degradation of methyl orange, Sep. Pur. Technol. 108 (2013) 159–165. [3] R. Andreozzi, V. Caprio, A. Insola, R. Marotta, Advanced oxidation processes (AOP) for water purification and recovery, Catal. Today 53 (1999) 51–59. [4] S. Yang, P. Wang, X. Yang, L. Shan, W. Zhang, X. Shao, R. Niu, Degradation efficiencies of azo dye Acid Orange 7 by the interaction of heat, UV and anions with common oxidants: persulfate, peroxymonosulfate and hydrogen peroxide, J. Hazard. Mater. 179 (2010) 552–558. [5] P. Hu, M. Long, Cobalt-catalyzed sulfate radical-based advanced oxidation: a review on heterogeneous catalysts and applications, Appl. Catal., B 181 (2016) 103–117. [6] Y.F. Huang, Y.H. Huang, Identification of produced powerful radicals involved in the mineralization of bisphenol A using a novel UV-Na2S2O8/H2O2-Fe(II, III) twostage oxidation process, J. Hazard. Mater. 162 (2009) 1211–1216. [7] Y.H. Huang, Y.F. Huang, C.I. Huang, C.Y. Chen, Efficient decolorization of azo dye reactive black B involving aromatic fragment degradation in buffered Co2+/PMS oxidative processes with a ppb level dosage of Co2+–catalyst, J. Hazard. Mater. 170 (2009) 1110–1118. [8] G.P. Anipsitakis, D.D. Dionysiou, Degradation of organic contaminants in water with sulfate radicals generated by the conjunction of peroxymonosulfate with cobalt, Environ. Sci. Technol. 37 (2003) 4790–4797. [9] J. Sun, X. Li, J. Feng, X. Tian, Oxone/Co2+ oxidation as an advanced oxidation process: comparison with traditional Fenton oxidation for treatment of landfill leachate, Water Res. 43 (2009) 4363–4369. [10] R.H. Waldemer, P.G. Tratnyek, R.L. Johnson, J.T. Nurmi, Oxidation of chlorinated ethenes by heat-activated persulfate: kinetics and products, Environ. Sci. Technol. 41 (2007) 1010–1015. [11] G.D. Fang, D.D. Dionysiou, Y. Wang, S.R. Al-Abed, D.M. Zhou, Sulfate radical-based degradation of polychlorinated biphenyls: effects of chloride ion and reaction kinetics, J. Hazard. Mater. 227 (2012) 394–401. [12] M. Nie, Y. Yang, Z. Zhang, C. Yan, X. Wang, H. Li, W. Dong, Degradation of chloramphenicol by thermally activated persulfate in aqueous solution, Chem. Eng. J. 246 (2014) 373–382. [13] T. Zhang, Y. Chen, Y. Wang, J.L. Roux, Y. Yang, J.P. Croué, Efficient peroxydisulfate activation process not relying on sulfate radical generation for water pollutant degradation, Environ. Sci. Technol. 48 (2014) 5868–5875.

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