Determining the key factors of nonradical pathway in activation of persulfate by metal-biochar nanocomposites for bisphenol A degradation

Journal Pre-proofs Determining the key factors of nonradical pathway in activation of persulfate by metal-biochar nanocomposites for bisphenol A degradation Haoyu Luo, Qintie Lin, Xiaofeng Zhang, Zhuofan Huang, Hengyi Fu, Rongbo Xiao, Shuang-shuang Liu PII: DOI: Reference:

S1385-8947(19)32970-5 https://doi.org/10.1016/j.cej.2019.123555 CEJ 123555

To appear in:

Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

3 July 2019 17 October 2019 18 November 2019

Please cite this article as: H. Luo, Q. Lin, X. Zhang, Z. Huang, H. Fu, R. Xiao, S-s. Liu, Determining the key factors of nonradical pathway in activation of persulfate by metal-biochar nanocomposites for bisphenol A degradation, Chemical Engineering Journal (2019), doi: https://doi.org/10.1016/j.cej.2019.123555

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Determining the key factors of nonradical pathway in activation of persulfate by metal-biochar nanocomposites for bisphenol A degradation

Haoyu Luo, Qintie Lin*, Xiaofeng Zhang, Zhuofan Huang, Hengyi Fu, Rongbo Xiao, Shuang-shuang Liu* Guangdong Industrial Contaminated Site Remediation Technology and Equipment Engineering Research Center, School of Environmental Science and Engineering, Guangdong University of Technology, Guangzhou 510006, China

Corresponding Authors: Qintie Lin; E-mail: [email protected]; Shuang-shuang Liu; E-mail: [email protected]

1

Abstract: In this study, three types of metal-biochar nanocomposites (CuO/BC, Fe3O4/BC and ZnO/BC) were synthesized as the catalysts to activate sodium persulfate (PS) for the degradation of bisphenol A (BPA). The results showed that nonradical pathway was the dominant reaction which was based on electron transfer intermediates in the CuO/BC-PS system (kobs ca. 0.0607 min-1) that consumed a small amount of PS (0.17 mM) for the mineralization of all BPA. Meanwhile, the higher kobs would be obtained when contained scavengers, verifying that the generation of intermediates were accompanied by the side reaction (CuO/BC reacted with PS and water to generate •OH) in the CuO/BC-PS system. However, •OH and SO4•- were the dominant radicals in the Fe3O4/BC-PS system (without scavengers, kobs ca. 0.0037 min-1). Besides, •OH was the main radical in the ZnO/BC-PS system (without scavengers, kobs ca. 0.0046 min-1). The mechanism was summarized below: PS effectively bond to unsaturated bonds (i.e., C=O lactones) and aromatic structures in the nanocomposites for the generation of the electron transfer intermediates, in which the stabilities of intermediates and the electron transfer capacities of nanocomposites themselves played the critical roles. In addition, nonradical degradation pathway of BPA was proposed, and could give a new insight into metal-carbon nanocomposites activating PS via nonradical pathway.

Keywords: Nonradical; Electron transfer; Persulfate; CuO; Biochar

1. Introduction The advanced oxidation technology (AOT) provides a convenient method for the degradation of refractory pollutants and has been widely applied to the remediation of contaminated soil[1], 2

groundwater[2] and wastewaters[3]. Recent studies have shown that the nonradical pathway can efficiently degrade refractory pollutants without the reliance on free radicals (•OH and SO4•-)[4, 5], which has greatly enhanced the AOT application in remediation of complex environments (for instance, interfering ion[6], alcohols[7] and natural organic matter[8]). The carbon-based catalysts, such as carbon nanotube[9], graphitic[10], nanodiamonds[11], have many chemical properties including the sp2/sp3 structures, the degree graphitization, 1O2 and the doped nitrogen, which facilitate their combination with peroxymonosulfate (PMS)[12]. Therefore, more demanding conditions are required in the process of preparation of the catalysis in order to meet the structural requirements[5]. These structures are crucial for the nonradical pathway. For example, the high graphitization, rich surface N-doped active sites and porous structure play key roles in the nonradical pathway for the carbon-based catalysts[13, 14]. It has been shown that N-graphene could be used as a catalyst to activate PMS to generate 1O2, which could degrade organic pollutants via nonradical pathway[15]. Nevertheless, the electrons transferred from phenol to PMS are probably responsible for the decomposition of pollutants via nonradical pathway[9]. The characteristics of the catalysts are particularly important because they may determine the reaction pathways. However, relatively less has been explored on the key role of crystal structures and functional groups in nonradical pathway, especially the interaction between metals and non-metals. Therefore, exploring the role of the active species in the catalysts to activate PS through the nonradical pathway for the degradation of pollutants is of critical importance. PMS, due to its dissymmetrical structure, can readily combine with the carbon catalysts and has been widely used in nonradical pathway as an electron acceptor. Compared with PMS, sodium persulfate (PS) is stable, cheap and easy to be preserved[4]. However, PS with symmetrical 3

structure is difficult to be activated by metal-free catalysts (i.e., carbon catalysts) in the nonradical pathway[9]. PS is often activated by the functional groups of carbon-based catalysts in radical pathway[16], implying the importance of the functional groups in the activation pathways. However, it remains unclear whether PS can undergo nonradical pathway degradation of pollutants through metal-carbon catalysts activation. Meanwhile, the high-temperature condition (>700 oC) and complex production process of carbon-based catalysts restrict their applications in catalysis, while metal catalysts can be synthesized under low-temperature condition and widely used in various catalytic reactions[17]. Unfortunately, the dissolution of metal ions and the low activity under alkaline condition of metal catalysts limit their applications. Therefore, research on the development of new metal-biochar catalysts with high activity and high stability for the nonradical activation of PS and exploring its activating mechanism are warranted. Three metal oxides were selected because the relative molecular weight and the proportion of oxygen element were similar. (i) CuO, a typical transition metal with strong electron transfer capacity and a p-type semiconductor[18], is common and cheap, and applied in many fields, such as catalysis[19], supercapacitor[20], environmental remediation[21]; (ii) Fe3O4, a classical heterogeneous catalyst (radical pathway), is widely studied and applied in environmental remediation[22]; (iii) ZnO, a semiconductor material with many free-moving electrons, is applied in the field of photocatalysis[23], supercapacitor[24], transition metal catalysis[25]. Loading three metal oxides onto the biochar to study the interaction between them on the catalytic activity. In addition, BPA, a typical environmental hormone[26] widely used as the chemical materials in various industries[27] which may have the potential adverse health effects[28], was selected as the target pollutant. In this study, the biochar, a waste solid carbon from a gasification plant, could be 4

transformed into treasure by the material preparation. We have been exploring the method that changed the waste biochar to the catalyst with high catalytic activity. The detail physical and chemical characteristics of this raw biochar could be found in our previous study[29]. Fourier Transform Infrared Spectroscopy (FTIR) and Boehm method were used to verify the combination between the functional groups and metal oxides. The nonradical degradation mechanism of BPA was verified by degradation kinetics, quenching experiments, X-ray Photoelectron Spectroscopy (XPS) and Electron Paramagnetic Resonance (EPR). Furthermore, the mechanism of electrons transfer was illustrated by electrochemical experiments.

2. Experimental Section 2.1 Materials and Chemicals Biochar was collected from a gasification plant (details on how to obtain the biochar and its characterization analyses could be found in our previous study[29]. Chemicals were shown in Supporting Information (SI) Text S1. 2.2 Synthesis and characterization of nanocomposites Biochar (less than 60 mesh) was first immersed in HNO3 (0.1 M) solution with a magnetic stirrer for 24 hours (for the removal of the metal foreign matters and the purpose of washing BC), then washed with deionized water for several times and dried at 60 oC in an oven before use. CuO/BC nanocomposites were synthesized using a hydrothermal method[30]. The preparation of the CuO/BC thorn-like nanocomposites was as follows: 2.50 g Cu(CH3COO)2•2H2O was dissolved in 80 mL deionized water. Then, 1.00 g BC and 1.25 g urea were added into the solution, then transferred to a 150 mL Teflon-lined stainless-steel autoclave, which then kept in 5

hydrothermal synthesis at 120 oC for 12 hours. After cooling to room temperature, the materials were separated by a suction filter, cleaned with ethanol and deionized water for several times, and then dried. Finally, the materials were synthesized in a vacuum tube furnace at 350 oC for 2 hours. ZnO/BC flower-like nanocomposites were synthesized using the same method (the urea precipitation method) except that 2.70 g Zn(CH3COO)2•2H2O and 1.48 g urea were added to the reaction system prior to the reaction. Fe3O4/BC sphere-like nanocomposites were synthesized using the improved method[31]. Approximately 3.46 g FeCl3•6H2O and 7.66 g CH3COONa were dissolved in 70 mL ethanediol. After 30 min, 1.00 g BC was added the solutions, then transferred to a 150 mL Teflon-lined stainless-steel autoclave, which then kept in hydrothermal synthesis at 200 oC for 8 hours. After cooling to room temperature, the materials were separated by a suction filter, cleaned with ethanol and deionized water for several times, and then freeze-dried for further use. The growth processes of three nanocomposites were illustrated in Scheme 1. The morphology and particle size of nanocomposites were analyzed by a SU8220 (Hitachi, Japan) Field emission scanning electron microscopy (FE-SEM). The porosity and specific surface area were calculated from BET method using an ASAP 2460 surface area and porosity analyzer (Micromeritics, USA). Internal properties of nanocomposites were measured by Field emission transmission electron microscope (FE-TEM) using a Talos F200S system (FEI, Czech). The functional groups of nanocomposites were analyzed by a Nicolet 6700 FTIR spectrometer (Thermo Scientific, USA). The crystal structures of nanocomposites were examined by Single crystal X-ray diffraction (XRD) using an D8 VENTURE X-ray diffractometer (Bruker, Germany). Elements valence-state change on nanocomposite’s surface were analyzed by XPS using an 6

Escalab 250Xi (Thermo Fisher, UK) 2.3 Experimental procedures The catalytic performance of the nanocomposites was investigated by activating PS for the degradation of BPA. All experiments were performed in 250 mL conical flasks containing 100 mL of the target pollutant solutions. In a typical experiment, different types of nanocomposites were added in conical flasks, which contained 100 mL BPA solutions (0.10 mM). First, 1 mM PS was dispersed in the initial reaction solutions, then the conical flasks were placed in the constant temperature oscillator at 155 rpm at 25 oC. At specific intervals, the reaction solutions were extracted and filtered through the 0.45 μm poly tetra fluoroethylene (PTFE) membranes, then mixed with equal volume of ethanol to quench the reactions, and finally the samples were stored in 4 oC for further analysis. The samples were measured by a Waters e2695 Alliance High performance liquid chromatography (HPLC) (Waters, USA) with an ultraviolet detector (Waters 2998 PDA) and a C-18 column. Acetonitrile (50%) and methanol (50%) were used as mobile phase and the velocity was set at 0.6 mL min-1 for BPA measurement (277 nm)[32]. All samples were performed in triplicate, and the results were displayed through the y-error bars. The Boehm method[33, 34] was used to measure the functional groups. The concentration of PS was measured by iodimetry[35]. The electron transfer capacity of nanocomposites was estimated by Cyclic voltammetry (CV), Electrochemical impedance spectroscopy (EIS)[7, 20]. The free radicals were determined by EPR experiments[29]. Total organic carbon (TOC) was employed to verify the mineralization abilities of three systems. Nonradical degradation pathway of BPA was verified by an Ultra performance liquid chromatography (UPLC)-Mass spectrum (MS)/MS (Waters Xevo TQ-S, USA). Details of this section were shown in SI Text S1. 7

3. Results and Discussion 3.1 The physical properties of nanocomposites BC was rich in pore structures under SEM (SI Fig. S1), which were probably to enhance material growth and provide more active sites[13]. CuO/BC exhibited abundant thorn-like structures (i.e., CuO) that were grown from BC surface and occupied the pore structures of BC (Fig. 1a and 1d). Fe3O4/BC formed chains of sphere-like structure, which was probably due to the magnetic character of Fe3O4 (Fig. 1b and 1e). ZnO/BC produced flower-like structures, which were formed by the sheet-shaped ZnO (Fig. 1c and 1f). Combined with BET technique, BC was typically mesoporous in structure belonging to “Type IV” which had an “H1” hysteretic effect (SI Fig. S2). This indicated that BC had numerous cylindrical holes (a uniform distribution of large holes), as evidenced from the SEM images. CuO/BC belonged to “Type IV” as well, but it showed an “H3” hysteretic effect (SI Fig. S3)[36], implying the irregular pore structures of CuO/BC. The average grain diameter of CuO nanoparticles was approximately 34.3 nm and nano-CuO thorn-structures (1.20 um × 130.00 nm) were formed by the agglomeration of CuO nanoparticles (SI Fig. S4). The structural characteristics of Fe3O4/BC were “Type IV” and showed an “H1” hysteretic effect (SI Fig. S5), suggesting that the original structure of BC was not altered after the loading of Fe3O4, and the combination of Fe3O4 with BC occurred only on the surface of Fe3O4/BC. Fe3O4 was evenly distributed on BC surface with an average grain diameter of approximately 315.00 nm (SI Fig. S6). The BET characteristics of ZnO/BC was similar to that of CuO/BC (SI Fig. S7). However, ZnO nanoparticles were grain-shaped with a diameter of approximately 22.80 nm, while ZnO nanoparticles were sheet-shaped with a thickness of 8

approximately 40.0 nm (SI Fig. S8). In addition, the flower-like ZnO nanostructures were formed on BC surface. 3.2 The crystal structures and growth mechanism The nanocomposites were synthesized using the same hydrothermal method. The crystallinity degree of CuO/BC (Fig. 2a) was 14.27% (the mass fraction ratio of the thorn-like CuO was 49.20%), and the peaks at 32.5°, 35.5°, 38.7°, 48.7°, 53.5°, 58.3°, 61.5°, 66.2°, 68.1°, 72.4° and 75.2° belonged to the monoclinic system, which were corresponding to the (110), (-111), (111), (-202), (020), (202), (-113), (022), (220), (-312) and (203) planes, respectively (PDF 00-005-0661)[37]. The new crystal structure appeared at 29.2°, which could be caused by the interaction between CuO and the functional groups. The crystallinity degree of Fe3O4/BC (Fig. 2b) was zero (the mass ratio of the sphere-like Fe3O4 was 36.5%), and peaks at 18.3°, 29.4°, 35.5°, 43.1°, 53.5°, 57.1° and 62.8° were ascribed to the cubic system, which were corresponding to the (111), (220), (311), (400), (422), (511) and (440) planes, respectively (PDF 00-003-0863)[38]. The crystallinity degree of ZnO/BC (Fig. 2c) was 39.12% (the mass ratio of the flower-like ZnO was 46.90%), and the peaks at 31.6°, 34.3°, 36.2°, 47.4°, 56.6°,62.7°, 66.3°, 67.8°, 69.1° and 76.9° were ascribed to the hexagonal system, which were corresponding to the (100), (002), (101), (102), (110), (103), (200), (112), (201) and (202) planes, respectively (PDF 00-005-0664)[39]. A new crystal structure generated at 29.8°, which might be due to the combined reaction of ZnO and BC. The possible growth mechanisms were shown in Scheme 1. The thorn-like structures were formed on the BC surface due to the growth of the (-111), (111) and (110) planes of CuO nanoparticles (Fig. 1g) and a monoclinic thorn-like CuO/BC was finalized after calcination. The high stable surfaces of CuO were (111) (ca. 0.74 J m-2) and (-111) (ca. 0.86 J m-2) which had the 9

low surface energy[18], indicating that the CuO/BC would have a high crystallinity degree and a high stability. The Fe3O4 nanoparticles were directly formed on the active sites of BC due to the growth of the (311) and (511) planes (Fig. 1h), which spread on the BC surface. But because of its magnetic, accumulated Fe3O4 would form a chain-like structure. The main difference between CuO/BC and ZnO/BC was that single crystal sheets were formed on BC surface due to the growth of (002) and (101) planes of ZnO nanoparticles (Fig. 1i) and the single crystal sheets were intertwined to form the polycrystalline flower-like structures after calcination. CuO and ZnO with higher crystallinity degree had better bonding with BC during growth, therefore, new crystal structures were produced. Three nanocomposites were stable under neutral condition (initial pH = 7 without the buffer solution), and the crystal structures were still obvious after reaction (Fig. 2a-c). Moreover, approximately 1.80 mg L-1 of Cu2+, 0 mg L-1 of Fe3+ and 1.36 mg L-1 of Zn2+ were in their reaction solutions after180 min (the finial pH = 5.8) (Fig. 3a). In addition, comparing BET between fresh and after 180 min reaction could also verify the stability of nanocomposites. BET only increased a little bit for the three nanocomposites after 180 min reaction (CuO/BC from 286.23 to 289.70 m2 g-1) (Fig. S9) (Fe3O4/BC from 206.25 to 212.72 m2 g-1) (Fig. S10) (ZnO/BC from 258.41 to 271.65 m2 g-1) (Fig. S11). It might be that the pores of BC increased slightly, but all of the three nanocomposites could keep a high stability during the reactions. The dissolution of metal ions indicated that the structural stability of nanocomposites was strong during the degradation process. 3.3 Degradation kinetics BPA was degraded by approximately 80.47 and 100% at 0.2 mM PS and more than 1 mM PS in 180 min, respectively (SI Fig. S12). This suggested that increasing the PS dosage would 10

promote its combination with the active sites on CuO/BC (0.4 g L-1) to generate intermediates, which were responsible for the degradation of BPA. However, the kobs was maintained relatively stable at approximately 0.0607 min-1 at the concentration of PS greater than 1 mM (Fig. 3b). Since the self-quenching reactions by excessive PS would not happen and it was therefore likely to conclude that the coupling reaction between CuO/BC and PS could generate the active intermediates to degrade BPA, indicating that kobs only depend on the amount of active intermediates[21]. Meanwhile, the Fe3O4/BC (0.4 g L-1)-PS system showed a degradation efficiency of 26.98 and 33.91% at 0.2 and 4 mM PS, respectively, within 180 min (SI Fig. S13). The kobs (ca. 0.0037 min-1) in the Fe3O4/BC-PS system was much lower than that in the CuO/BC-PS system (Fig. 3b). Furthermore, the ZnO/BC (0.4 g L-1)-PS system exhibited a degradation efficiency of 34.27 and 40.66% for the BPA degradation at 0.2 and 4 mM of PS within 180 min, respectively (SI Fig. S14). The kobs (ca. 0.0046 min-1) in the ZnO/BC-PS system was slightly higher than that in the Fe3O4/BC-PS system containing 0.4 g nanocomposite dosages and 1mM of PS. These results indicated that PS needed to reach a certain concentration to degrade all BPA in the CuO/BC-PS system. The effects of nanocomposite dosage on the degradation of BPA was also investigated (Fig. 3c). The kobs for BPA degradation in different systems ranged from 0.04 to 0.8 g L-1 for the three nanocomposites at different dosages at a fixed PS (1 mM), but it was much higher in the CuO/BC-PS system than that in the other systems. However, 0.04 g L-1 CuO/BC could also degrade all BPA at 1 mM PS within 180 min (SI Fig. S15). The main contribution to the increase in the kobs from 0.0096 to 0.0984 min-1 was CuO/BC dosage. These results indicated that the interaction between PS and the active sites of CuO/BC reached the equilibrium and the active 11

intermediates generated from the reaction were saturated[40]. Therefore, when arrived the lowest concentration of PS capable of degrading BPA, the degradation rate depended on the dosage of nanocomposites. In the Fe3O4/BC-PS and ZnO/BC-PS systems, the maximum BPA degradation efficiencies were 47.94 and 53.56%, respectively (SI Fig. S16 and S17). It showed a low kobs in the Fe3O4/BC-PS system (ca. 0.0037 min-1) and in the ZnO/BC-PS system (ca. 0.0046 min-1). The above results indicated that the activity of the systems was mainly determined by the nanocomposites, 0.4 g L-1 nanocomposites and 1 mM PS were selected as the dosage for the study of the degradation mechanism of BPA (Fig. 3d). Thus, compared to radical pathway activation of PS, nonradical pathway might be more dependent on the characteristics of the catalysts and combination reaction with PS. 3.4 Identify active intermediates and reaction pathways The Boehm method[33, 34] provides the bases for quantitative analysis of the functional groups of nanocomposites (Table 1). To eliminate potential errors, the CuO, Fe3O4, ZnO groups were as the blank experiments in Boehm experiments. The results showed that the functionals groups (i.e. carboxyl groups, hydroxyl groups and lactones) onto those metal oxides would be basically ignored (data not shown). The BC had the greatest number of functional groups, with the carboxyl groups (ca. 1.45 mmol g-1) being the dominant one. The functional groups on the surface of BC were altered significantly after the loading of the metal oxides because the reaction between CuO and the functional groups would convert the other two functional groups to the lactones (ca. 1.065 mmol g-1), which was occurred similarly for ZnO/BC. However, the combination degree of ZnO/BC (ca. 0.81 mmol g-1 of lactones) was less than that of CuO/BC (ca. 1.065 mmol g-1 of lactones). The difference in the combination degree could be explained by the thorn-like structure 12

of CuO, which had a large contact area and might accelerate the binding reaction to the functional groups on BC. Therefore, CuO/BC had more active sites than ZnO. Nevertheless, the characteristics of sphere-like Fe3O4/BC was completely different from those of CuO/BC and ZnO/BC in that the carboxyl groups were converted to hydroxyl groups (ca. 0.475 mmol g-1) and lactones (ca. 0.60 mmol g-1) by Fe3O4. This meant that the interaction between BC and Fe3O4 was weaker than that between BC and CuO or ZnO, resulting in much fewer active sites (i.e., lactones) on the chain of the Fe3O4/BC. FTIR spectra of the three nanocomposites were not obviously different in the 0 min and all metal oxides were combined with the functional groups on the BC surface (Fig. 2d-f). On the other hand, the lactones and aromatic structures in the nanocomposites combined with S=O=S and sulfonate groups[8, 40], which could generate the active intermediates for electrons transfer and resulting in the increase of the peak height. As to the CuO/BC-PS system, lactones and aromatic structures would be combined with PS to generate active intermediates within the first 30 min, and they were stable even after 180 min (Fig. 2d), while in the Fe3O4/BC-PS system, the active intermediates were unstable, and could be consumed by heterogeneous activation to generate free radicals (Fig. 2e). For the ZnO/BC system, the increasing peaks at lactones and aromatic structures indicated that the intermediates were consumed by single electron transfer process to generate free radicals (Fig. 2f). The Cu 2p spectra shown in Fig. 4ab, the valence-state change of Cu was not obvious in the reaction (Cu 2p3/2 from 45.38 to 48.65%). This indicated that the activation of PS by CuO/BC was not dependent on single electron transfer of metal. For Fe 2p spectra (Fig. 4cd), the peaks at 709.80 (Fe2+) and 712.98 (Fe3+) changed from 45.15% and 15.08% to 38.66% and 17.01% after 13

reaction, respectively[41], indicating that the valence-state change of Fe (from Fe2+ to Fe3+) was the mainly activation pathway in Fe3O4/BC-PS system. For Zn 2p spectra (Fig. 4ef), the peaks at Zn 2p3/2 (from 63.64 to 63.48%) and Zn 2p1/2 (from 36.36 to 36.52%) hardly changed after reaction, indicating that free-moving electrons was the main reason for the activation of PS. The O 1s spectra showed (Fig. 5) that the peaks at 528.97, 530.76 and 532.97 belonged to O2-, C-O and C=O, respectively[29]. After reaction, the O2- and C-O in CuO/BC decreased from 34.09% and 53.45% to 7.71% and 39.41%, respectively. This was obviously different from the other systems, which meant that the nonradical pathway was the dominant pathway in the CuO/BC-PS system (Fig. 5ab). Because the O2- and C=O in conventional Fe3O4/BC catalyst decreased after reaction, with the C=O changing from 28.73 to 7.71% particularly (Fig. 5cd). The ZnO/BC-PS system (Fig. 5ef) had a similar trend as the Fe3O4/BC-PS system, but the difference was that O2- decreased significantly (from 66.66 to 37.18%). It could be concluded that the free-moving electrons in ZnO/BC would activate PS to generate free radical[25]. For nonradical pathway, the formation of active intermediates might be related to the unsaturated bonds (i.e., C=O, lactones) which could bond with PS to generate electron transfer complex[9, 21]. 3.5 Activity of nonradical pathway The quenching experiments illustrated the activity of nonradical pathway. The scavengers (100 mM) had no effect on the BPA degradation, so did the coexistence of PS and scavengers (SI Fig. S18). The nonradical degradation of BPA was not inhibited by the free radical scavengers in the CuO/BC-PS system (Fig. 6a), and the kobs were approximately 0.0610 min-1 (10 mM tert butyl alcohol (TBA), k = 3.8-7.6 × 108 M-1 S-1 for •OH[42]), 0.0597 min-1 (100 mM TBA), 0.0665 min-1 (10 mM ethanol, k = 1.6-7.7 × 107 M-1 S-1 for SO4•- and k = 1.2-2.8 × 109 for •OH[43]) and 0.0666 14

min-1 (100 mM ethanol). This directly proved that the CuO/BC-PS system could completely degrade all BPA without free radicals, and maintained a high degradation rate. On the other hand, L-histidine was selected as a scavenger (k = 3.2 × 107 M-1 S-1[42]) to exclude the effects of 1O2 on BPA degradation. The results showed that the concentrations of L-histidine being 10 mM and 100 mM, the kobs could remain at approximately 0.0674 and 0.0660 min-1, respectively. PS activated by 1O2 in nonradical pathway was not the main reaction, thus demonstrating the aforementioned conclusion that the formation of electron transfer complex might be the dominant active species in the CuO/BC-PS system[9, 21]. To elucidate radical pathway in the Fe3O4/BC-PS system, the addition of scavengers on BPA degradation was investigated (Fig. 6b). When the 100 mM ethanol was added, the radical pathway was fully inhibited in the Fe3O4/BC-PS system, showing a similar BPA degradation efficiency (11.45%) to adsorption (ca. 11.05%) (SI Fig. S19). The degradation efficiency of BPA when adding 10 mM TBA was greater than adding 10 mM ethanol, indicating that the radical pathway producing •OH and SO4•- was the dominant reactions in the Fe3O4/BC-PS system[38]. The competitive degradation between BPA and L-histidine occurred due to the indiscriminate attack of free radicals, and the inhibition would be obvious when adding the more amount of L-histidine. The ZnO/BC-PS system and the Fe3O4/BC-PS system were similar in radical pathway, and the degradation of BPA was also inhibited by alcohols, but the inhibition was similar when adding 10 mM TBA (ca. 29.41%) and 10 mM ethanol (ca. 29.56%), indicating that •OH was the main radical in the ZnO/BC-PS system (Fig. 6c). The degradation efficiency when adding 100 mM ethanol (ca. 24.96%) was slightly greater than ZnO/BC adsorption (ca. 24.85%) and BC adsorption (ca. 16.48%). This clarified that it had a weak nonradical pathway in the ZnO/BC systems for the degradation of BPA. Meanwhile, the degradation of BPA was markedly 15

inhibited by 100 mM L-histidine (ca. 26.56%) due to the indiscriminate attack of free radicals. The activity of nonradical pathway might be determined by the electron transfer capacity of nanocomposites themselves as the adsorption of BPA by CuO/BC was much higher than by any others (SI Fig. S19). The nonradical pathway in the CuO/BC-PS system did not rely on free radicals to degrade pollutants, and consumed a small concentration of PS (ca. 0.17 mM) to degrade all BPA (Fig. 6d), which greatly improved the utilization of oxidant. A small amount of active site was generated via the combination of sphere-like Fe3O4 with BC, and approximately 0.28 mM of PS was activated to generate free radicals which would lead to a poor degradation efficiency of BPA. The free-moving electrons in ZnO/BC would activate PS to generate free radicals, and then the intermediate would be destroyed, resulting that the activity of nonradical pathway was weak[25]. In addition, the ratio of PS to BPA removal were selected as specific oxidant efficiency (SOE)[44], the units of this equation were mM (Eq. (1) ). The results indicated that the highest SOE (ca. 58.1%) was obtained in the CuO/BC-PS system when compared with that in the Fe3O4/BC-PS (ca. 11.5%) and the ZnO/BC-PS (ca. 12.5%) systems. It meant that the nonradical pathway could obtain a higher oxidant efficiency in BPA degradation process. SOE =

BPA0 - BPAt

(1)

PS0 - PSt

3.6 Mechanisms of free radicals and electrons transfer EPR spectra revealed the radical mechanism for the degradation of BPA. The PS system had low-intensities of DMPO-OH and DMPO-SO4- (Fig. 7a), which might be generated by the hydrolysis of PS[45]. For the BC-PS system, it had slightly larger intensities due to activation of PS by the functional groups on BC surface[29]. For the Fe3O4/BC and ZnO/BC, they activated PS 16

via a radical pathway to generate •OH and SO4•- with high-intensities. These results proved that the radical pathway was the dominant reaction pathway in those systems for the degradation of BPA. The high DMPO-OH intensity and only those specific peaks (1:2:2:1 for DMPO-OH) were observed in the CuO/BC-PS system, but the quenching experiments showed that BPA was not quenched by scavengers, meaning that BPA was mainly degraded by nonradical pathway in this system, inferring •OH might be generated by the generation process of the intermediates (i.e. side reaction) in the CuO/BC-PS system (the detail information of side reactions showed in Graphical abstract). Another evidence, the capture experiments in dimethyl sulfoxide (DMSO), could confirm that whether PS was activated to generate SO4•- directly or not. The results showed that no intensity in the CuO/BC-PS system was observed. It could be inferred that CuO/BC as a catalyst would activate PS to generate •OH via reaction with water and PS[29, 46]. In addition, the concentration of PS at different time in the CuO/BC-PS system could also confirm that the fast consumption process of PS was observed within the first 10 min, indicating that the generation of intermediates were accompanied with the side reactions until the formation of intermediates were completed (after 10 min). Then PS was no longer consumed after 10 min as the generation reaction of intermediates had been completed (Fig. 6d). Quenching experiments in the CuO/BC-PS system also verified that the generation of •OH resulted from the side reaction, because kobs in the system containing scavengers were higher than that in the systems without scavengers (Fig. 6a). In conclusion, these results provided direct evidence on the degradation of BPA by the nonradical pathway in the CuO/BC-PS system. Ho et al.[47] illustrated that the N-doping sites on N-doped biochar surface could act as an electron center to transfer electron, enhancing the pollutants degradation by nonradical pathway. 17

Yun et al.[9] explained that nanotube could combine with PMS to generate electron transfer complex that could transfer electron from pollutants to PMS. Previous studies showed that the electron transfer ability of carbon-based materials was crucial for nonradical pathway. So, those abilities of metal-biochar nanocomposites must be considered. The results showed that the integral area of CuO/BC was the largest among the four groups at the 100 mV s-1 scan rate in CV curves, followed by ZnO/BC, Fe3O4/BC and BC (Fig. 7b), implying CuO/BC had the largest capacitance property, which could transfer and store more electrons, therefore higher catalytic activity in oxidation-reduction reaction. Thorn-like structure signified that CuO/BC would have more active sites and high specific surface area. In addition, the electron transfer capacity of CuO/BC increased rapidly as the functional groups were transformed to lactones. These characteristics would enhance the electron transfer between metals and nonmetals[18]. However, the redox peak was observed only in the CuO/BC system. CuO would become an electronic center when combined with BC[19, 46] and electrons could assemble on the active sites of CuO/BC surface for a high-performance in catalysis[20]. For Fe3O4/BC, it exhibited ordinary capacitance property. Therefore, sphere-like structures may lead to low bonding strength when Fe3O4 was combined with BC, thereby increasing the energy barrier for electron transfer. Lactones in Fe3O4/BC was not sufficient to transfer electrons effectively and electrons could be difficult to freely shuttle due to the magnetic force of Fe3O4, and the intermediates were destroy by Fe3O4/BC activated PS[41]. As for ZnO/BC, it showed a high capacitance property due to the relative higher bonding strength of the flower-like ZnO with BC. It was noteworthy that there was an oxidation peak in the first cycle when ZnO/BC was added, and the peak would disappear after several cycles (data not shown). This suggested that there were some free-moving electrons in the freshly prepared 18

ZnO/BC, PS was mainly activated by free-moving electrons from ZnO/BC, then the intermediates would be destroyed. But the oxidation-reduction reaction could not form a cycle due to the one-way movement of the electrons. The extent of difficulty to which the electron transferred could be explained by EIS curves[48]. In the EIS curve for CuO/BC (Fig. 7c), the maximum slope indicated that it was most ready for the electrons to be transferred between CuO/BC and solutions relative to the other two nanocomposites. This further confirmed our early speculation that nonradical pathway was based on electron transfer process, in which the characteristics of the nanocomposites played a decisive role[21]. The EIS curve for Fe3O4/BC had a minimum slope, indicating the electrons could not be transferred freely between Fe3O4/BC and solutions due to magnetic. The electron transfer capacity of Fe3O4/BC was lower than BC. Therefore, the radical pathway and the electrostatic adsorption were dominant in Fe3O4/BC-PS system[22, 49]. The EIS curve for ZnO/BC had the second highest slope, implying a relatively easy electron transfer. However, a redox reaction did not occur spontaneously which led to a weak system for BPA degradation via nonradical pathways to activate PS. Furthermore, the reuse of nanocomposites (Fig. 7d) showed that CuO/BC had an excellent characteristic with high electron transfer capacity, low electrical resistance and stable crystal structure, resulting that the unsaturated bonds (i.e., C=O, lactones) combined with PS to produce the intermediates with high stability and activity, which could rapidly transfer electrons for the nonradical degradation of BPA. The degradation efficiency of the Fe3O4/BC-PS system gradually decreased with reuse times increasing, probably due to the consumption of active sites (Fe2+ in Fe 2p). After the first cycle, the main cause of BPA degradation efficiency decreasing in the 19

ZnO/BC-PS system was that the free-moving electrons would be consumed by the radical pathway and then no new ones had been added. 3.7 Mineralization ability of the systems and the degradation process of nonradical pathway The mineralization ability of three systems was also considered using TOC. Meanwhile, TOC from BC was removed when calculating TOC value (ca. 3.04 mg L-1). The results showed that the mineralization rates of BPA in the CuO/BC-PS, Fe3O4/BC-PS and ZnO/BC-PS systems were 100%, 23.51% and 34.96% after 180 min reaction, respectively (SI Fig. S20). This reflected that nonradical degradation of BPA could achieve a much higher mineralization rate. Nonradical degradation pathway of BPA in CuO/BC-PS were verified (Fig. 8), and aromatic compounds (2,4-di-tert-butyl-5-methylphenol, ethane-1,1-diyldibenzene and prop-1-en-2-ylbenzene), alcohols (cyclohexa-2,5-dien-1-ol and cyclohexanol) and acids (penta-1,4-dien-3-one, formic acid and maleic acid) were the main possible intermediate products[50-52]. The detail spectrums of UPLC-MS/MS were shown in SI Fig. S21. From intermediate products could find that there were less hydroxylation products in the reaction. Thus, the nonradical degradation pathway that was based on electron transfer to attack aromatic rings would be verified, and the CuO/BC-PS system could achieve 100% mineralization rate of BPA. 4. Conclusions This study demonstrated the key roles of metals and intermediates in activating PS via a nonradical pathway for BPA degradation. On one hand, CuO/BC had highest specific surface area and the largest capacitance property, resulting in the most active sites when compared with other nanocomposites. Thus, the nonradical pathway degradation of BPA with high kobs (ca. 0.0607 min-1) in the CuO/BC-PS systems would occur and was not inhibited by scavengers. On the other 20

hand, through quenching experiments, •OH and SO4•- were the main free radicals in the Fe3O4/BC-PS system, and •OH was the dominant free radical in the ZnO/BC-PS system. The nonradical pathway mechanism was that the intermediates were produced in the binding reactions of the nanocomposites with PS, and the lactones and aromatic structures in the nanocomposites were essential for those reactions. Meanwhile, the BPA degradation process was that transferring electrons from pollutants to PS via the electron transfer intermediates formed from the binding reactions between CuO/BC and PS. Therefore, there was a significant difference in the change of O-element between nonradical pathway and radical pathway, because intermediates generated due to unsaturated bonds (i.e., C=O, lactones) that promoted the combination between nanocomposites and PS. In addition, it could achieve a high mineralization rate of BPA (100%) in the CuO/BC-PS system, and the high activity for the degradation of BPA after 5 reuse times of CuO/BC was also observed. Moreover, our study provided a rapid degradation method of aromatic pollutants through nonradical pathway, as well as developed a simple and effective method to reuse factory solid waste (biochar).

Appendix A. Supplementary data Supplementary data associated with this manuscript can be found in Supporting Information (SI).

Acknowledgments This work was financially supported by National Key Research and Development Program of China (2018YFC1802803, 2017YFD0801302), National Natural Science Foundation of China (21677041, 41371317), Science and Technology Project of Guangdong Province 21

(2017B020216002), and YangFan Innovative and Entrepreneurial Research Team Project (2015YT02N012).

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25

BET surface area (m2 g-1)

Average pore radius (nm)

Volume of pore (cm3 g-1)

Carboxyl Groups (mmol g-1)

Hydroxyl groups (mmol g-1)

Lactones (mmol g-1)

Total groups (mmol g-1)

BC

431.38

2.21

0.087

1.450

0.400

0.000

1.850

CuO/BC

286.23

3.92

0.121

0.385

0.195

1.065

1.645

Fe3O4/BC

206.25

2.37

0.071

0.450

0.475

0.600

1.525

ZnO/BC

258.41

4.70

0.159

0.540

0.160

0.810

1.510

Table 1. The physical properties of materials

26

Fig. 1. SEM and TEM images of (a), (d), (g) CuO/BC; (b), (e), (h) Fe3O4/BC; (c), (f), (i) ZnO/BC.

27

Fig. 2. Characterizations of nanocomposites in reaction systems were detected at 0 min, 30 min and 180 min. XRD and FTIR spectra of (a), (d) CuO/BC; (b), (e) 28

Fe3O4/BC and (c), (f) ZnO/BC.

Scheme1. The growth mechanisms and the formation process of lactones

29

Fig. 3. (a) Metal ions in the reaction solutions; pseudo-first-order rate constants for BPA degradation by (b) different concentration of PS and (c) different dosage of nanocomposites; (d) the degradation curves of BPA. Initial condition: [nanocomposites]0 = 0.4 g L-1; [PS]0 = 1 mM; [BPA]0 = 0.1 mM; [pH]0 = 7; T = 25 ℃.

30

Fig. 4. Cu 2p spectra for CuO/BC (a) before reaction (0 min), (b) after reaction (180 min); Fe 2p spectra for Fe3O4/BC (c) before reaction (0 min), (d) after reaction (180 min); Zn 2p spectra for ZnO/BC (e) before reaction (0 min) (f) after reaction (180 min).

31

Fig. 5. O 1s spectra for CuO/BC (a) before reaction (0 min) (b) after reaction (180 min); for Fe3O4/BC (c) before reaction (0 min) (d) after reaction (180 min); for ZnO/BC (e) before reaction (0 min) (f) after reaction (180 min). 32

Fig. 6. Quenching experiments in (a) the CuO/BC-PS system; (b) the Fe3O4/BC-PS system and (c) the ZnO/BC- PS system; (d) remaining concentration of PS. Initial condition: [nanocomposites]0 = 0.4 g L-1; [PS]0 = 1 mM; [BPA]0 = 0.1 mM; [pH]0 = 7; T = 25 ℃. 33

Fig. 7. (a) EPR spectra of DMPO-OH and DMPO-SO4- adducts produced by different systems in 2 mM phosphate buffer saline solution; (b) cyclic voltammetry curves (100 mV s-1 scan rate); (c) electrochemical impedance spectroscopy curves in neutral electrolyte (1M Na2SO4); (d) the reusability of nanocomposites for the degradation of 34

BPA. Initial condition: [nanocomposites]0 = 0.4 g L-1; [PS]0 = 1 mM; [BPA]0 = 0.1 mM; [pH]0 = 7.4; T = 25 ℃; ●= DMPO-OH; ▲= DMPO-SO4-.

Fig. 8. Proposed nonradical degradation pathway of BPA in the CuO/BC-PS system.

35