Cu/CuO-Decorated Peanut-Shell-Derived Biochar for the Efficient Degradation of Tetracycline via Peroxymonosulfate Activation

: Biochar (BC) usually has abundant surface functional groups, well-developed pore structures and high speciﬁc surface areas, which can combine with transition metals for peroxymonosulfate (PMS) activation to degrade organics. In this paper, BC modiﬁed with Cu/CuO was prepared by a modiﬁed impregnation pyrolysis method using peanut shells as raw materials. The morphology, structure and physicochemical properties were analyzed. Results showed that the originally smooth BC surface was modiﬁed into a rough structure with distributed metal particles, and the speciﬁc surface area of the modiﬁed Cu/CuO-BC700 (i.e., Cu/CuO-BC) increased from the initial 22.57 to 192.64 m 2 /g. The Cu/CuO-BC was employed for PMS activation and tetracycline (TC) degradation, achieving a removal efﬁciency of 93.2% at TC initial concentration 20 mg/L, PMS concentration 0.5 mM and catalyst dosage 0.1 g/L after 30 min. The inﬂuence of co-existing anions in the actual water on TC degradation followed the order of HCO 3 − > H 2 PO 4 − > Cl − , and HA had an inhibitory effect on TC degradation. A variety of active species participated in TC degradation, and the free radical pathway played a dominant role. Furthermore, the Cu/CuO-BC could maintain the degradation efﬁciency of TC up to 80% even after ﬁve consecutive cycles. The Cu/CuO-BC maintained high activity through redox reactions between catalytically generated active species and the cycling of metal ions (Cu + /Cu 2+ ).


Introduction
Tetracyclines (TCs) with strong bacterial inhibitory action are widely used for the prevention and treatment of human and animal diseases [1]. TCs, which are detected in surface water, groundwater and soil as an emerging contaminant through multiple migration pathways, cannot be fully metabolized in animals [2,3]. The relevant risk assessment shows that human daily intake of antibiotics (containing TCs) through drinking water is 1 µg/kg [4]. Even if the concentration of ng/L-µg/L of TCs exists in water for a long time, it still poses a great danger to human health and environmental security [5]. Therefore, the safe and highly effective removal of TCs is urgent.
Various methods, including biological treatment, adsorption, membrane filtration and advanced oxidation, are applied to remove TCs from water [6][7][8][9]. Biological methods have high treatment capacity and mild reaction conditions. Nevertheless, microorganisms in water are subject to massive death when treating TC-containing wastewater [10]. The adsorption process is effective, simple to operate and does not produce toxic substances during the removal of TCs [11]. In addition, the membrane filtration technology has the characteristics of low cost and simple operation. However, adsorption and membrane treatment methods could enrich antibiotics rather than degrade them, inducing a risk of secondary pollution [12]. Among them, advanced oxidation processes (AOPs) can oxidize TCs to small molecules or even achieve complete mineralization, which has a bright application prospect [13].
The surface morphology of the initial BC and Cu/CuO-BC composites were analyzed by SEM, as shown in Figure 1. The morphology of pure BC showed a porous structure with a relatively smooth outer surface. As seen in Figure 1b-d, the surface morphologies of Cu/CuO-BC500, Cu/CuO-BC600 and Cu/CuO-BC700 became rough, with fine particulate matter distributed on the surface and in the pore channels, which might be due to the conversion of metal ions into metal and oxides during the pyrolysis process. The mapping profile of Cu/CuO-BC700 is shown in Figure 1, which reveals the uniform distribution of Cu, C, O and N in the composite catalyst. The Cu element was more uniformly distributed on the surface of BC, which was beneficial to the catalytic activity.
The specific surface area of catalysts has an important effect on the activity, and in general, the larger the specific surface area, the higher the activity [28]. The specific surface area of BC was determined as 22.57 m 2 /g. Interestingly, the specific surface area of Cu/CuO-BC increased from 41.43 to 192.64 m 2 /g with the annealing temperature. Obviously, the pure BC had the smallest specific surface, and the specific surface area of Cu/CuO-BC700 was about 8.5 times larger than that of pure BC. More pyrolysis gas was generated at higher annealing temperatures, which facilitated the formation of pore structure [28]. As a result, the porous structure contributed to the increase of specific surface area. The pores of the catalyst after modification seemed more abundant from SEM images, which facilitated compound migration and removal. The specific surface area of catalysts has an important effect on the activity, and in general, the larger the specific surface area, the higher the activity [28]. The specific surface area of BC was determined as 22.57 m 2 /g. Interestingly, the specific surface area of Cu/CuO-BC increased from 41.43 to 192.64 m 2 /g with the annealing temperature. Obviously, the pure BC had the smallest specific surface, and the specific surface area of Cu/CuO-BC700 was about 8.5 times larger than that of pure BC. More pyrolysis gas was generated at higher annealing temperatures, which facilitated the formation of pore structure [28]. As a result, the porous structure contributed to the increase of specific surface area. The pores of the catalyst after modification seemed more abundant from SEM images, which facilitated compound migration and removal.

Phase Structure and Composition
The structural analysis of BC and Cu/CuO-BC700 was performed by X-ray diffractometer, as shown in Figure 2. The Cu/CuO-BC700 exhibited obvious diffraction peaks of Cu at 2θ = 43.39°, 50.43° and 74.13°, corresponding to the typical (111), (200) and (220) crystal planes (Cu PDF#04-0836) [29], respectively. Diffraction peaks of Cu were found at 35.41° and 38.70°, which were identified as typical (002) and (111) crystal planes of CuO (CuO PDF#48-1548) [30]. It can be inferred that the conversion of Cu 2+ in the impregnated solution to CuO and Cu monomers occurred during the high-temperature pyrolysis.

Phase Structure and Composition
The structural analysis of BC and Cu/CuO-BC700 was performed by X-ray diffractometer, as shown in Figure 2. The Cu/CuO-BC700 exhibited obvious diffraction peaks of Cu at 2θ = 43.39 • , 50.43 • and 74.13 • , corresponding to the typical (111), (200) and (220) crystal planes (Cu PDF#04-0836) [29], respectively. Diffraction peaks of Cu were found at 35.41 • and 38.70 • , which were identified as typical (002) and (111) crystal planes of CuO (CuO PDF#48-1548) [30]. It can be inferred that the conversion of Cu 2+ in the impregnated solution to CuO and Cu monomers occurred during the high-temperature pyrolysis. The specific surface area of catalysts has an important effect on the a general, the larger the specific surface area, the higher the activity [28]. The s area of BC was determined as 22.57 m 2 /g. Interestingly, the specific su Cu/CuO-BC increased from 41.43 to 192.64 m 2 /g with the annealing tempe ously, the pure BC had the smallest specific surface, and the specific su Cu/CuO-BC700 was about 8.5 times larger than that of pure BC. More pyro generated at higher annealing temperatures, which facilitated the formation ture [28]. As a result, the porous structure contributed to the increase of sp area. The pores of the catalyst after modification seemed more abundant f ages, which facilitated compound migration and removal.

Phase Structure and Composition
The structural analysis of BC and Cu/CuO-BC700 was performed by tometer, as shown in Figure 2. The Cu/CuO-BC700 exhibited obvious diffra Cu at 2θ = 43.39°, 50.43° and 74.13°, corresponding to the typical (111), (2 crystal planes (Cu PDF#04-0836) [29], respectively. Diffraction peaks of Cu 35.41° and 38.70°, which were identified as typical (002) and (111) crystal p (CuO PDF#48-1548) [30]. It can be inferred that the conversion of Cu 2+ in the solution to CuO and Cu monomers occurred during the high-temperature p   in surface functional groups, which contributed to the hydrophilicity and dispersibility of the catalyst. For Cu/CuO-BC700, the characteristic peak at 619 cm −1 in the low-frequency region was caused by the stretching vibration of the Cu-O bond [34], which proved the successful loading of Cu into the BC.
The FTIR spectra of the original BC and Cu/CuO-BC700 were examine sults are shown in Figure 3. For BC and Cu/CuO-BC700, the characteristic 1355, 1093, 805, 875 and 469 cm −1 in the mid-high and low-frequency band sented the C=C/C=O, C-H, C-O-C, C-H (1,2,4 substituted), aromatic C-H stretching vibration peaks on BC surface, respectively [31][32][33]. The original B surface functional groups, which contributed to the hydrophilicity and di the catalyst. For Cu/CuO-BC700, the characteristic peak at 619 cm −1 in the l region was caused by the stretching vibration of the Cu-O bond [34], whi successful loading of Cu into the BC.

The Effect of Pyrolysis Temperature
The effect of pyrolysis temperature on the activity of the catalysts was As shown in Figure 4a, the adsorption efficiency of TC on Cu/CuO-BC in 18.9% to 29.4% as the preparation temperature rose from 500 to 700 °C, wh tively correlated with the specific surface area. At the oxidative degradation concentration rapidly decreased in the first 10 min and then slowly decreas 30 min. The TC degradation efficiency was also increased with the rise of p perature (from 500 to 700 °C), increasing from 75.2% to 93.2% at 30 min. W (0.5 mM) was added, the removal rate of TC was 27.8% in 30 min, which active species produced by chemical bond breakage in PMS. To further i degradation kinetics of TC by PMS activation with Cu/CuO-modified BC, kinetic model was employed, and the fitting results are shown in Figure 4b all degradation data followed the first-order kinetic model (R 2 = 0.89-0.97). tion rate constant of MB by PMS was only 0.0115 min −1 , while that of Cu/CuO Cu/CuO-BC600/PMS and Cu/CuO-BC700/PMS systems were 0.0455, 0.05 min −1 , respectively. With the same TC/PMS concentration and lower cataly removal efficiency in the Cu/CuO-BC700/PMS system at 30 min was compar of the Fe/Mn-bimetallic-composite-modified biochar/PMS system at 60 min. tion rate constant by the Cu/CuO-BC700/PMS system was about twice as h the Fe/Mn-modified biochar/PMS system [35]. In summary, the Cu/CuO-BC to as Cu/CuO-BC) exhibited the desired activity for PMS activation and TC  The effect of pyrolysis temperature on the activity of the catalysts was investigated. As shown in Figure 4a, the adsorption efficiency of TC on Cu/CuO-BC increased from 18.9% to 29.4% as the preparation temperature rose from 500 to 700 • C, which was positively correlated with the specific surface area. At the oxidative degradation stage, the TC concentration rapidly decreased in the first 10 min and then slowly decreased from 10 to 30 min. The TC degradation efficiency was also increased with the rise of pyrolysis temperature (from 500 to 700 • C), increasing from 75.2% to 93.2% at 30 min. When only PMS (0.5 mM) was added, the removal rate of TC was 27.8% in 30 min, which relied on the active species produced by chemical bond breakage in PMS. To further investigate the degradation kinetics of TC by PMS activation with Cu/CuO-modified BC, the first-order kinetic model was employed, and the fitting results are shown in Figure 4b. Apparently, all degradation data followed the first-order kinetic model (R 2 = 0.89-0.97). The degradation rate constant of MB by PMS was only 0.0115 min −1 , while that of Cu/CuO-BC500/PMS, Cu/CuO-BC600/PMS and Cu/CuO-BC700/PMS systems were 0.0455, 0.0519 and 0.0921 min −1 , respectively. With the same TC/PMS concentration and lower catalyst dosage, the removal efficiency in the Cu/CuO-BC700/PMS system at 30 min was comparable with that of the Fe/Mn-bimetallic-composite-modified biochar/PMS system at 60 min. The degradation rate constant by the Cu/CuO-BC700/PMS system was about twice as high as that of the Fe/Mn-modified biochar/PMS system [35]. In summary, the Cu/CuO-BC700 (referred to as Cu/CuO-BC) exhibited the desired activity for PMS activation and TC degradation.

The Recyclability of Cu/CuO-BC
The reusability of the catalyst is closely related to practical applicability. The recyclability of Cu/CuO-BC was evaluated by a recycling experiment. As shown in Figure 5, the TC degradation efficiency was reduced from 93.2% to 83.5% after using it five times. The results indicated that the Cu/CuO-BC prepared in this study exhibited relatively good reusability. In addition, the leaching amount of metal ions in the solution was detected by ICP-MS. The results showed that the leaching amount of Cu 2+ was 0.22 mg/L, much lower than the Chinese Sanitary Standard for Domestic Drinking Water GB 5749-2022 [36]. After five consecutive cycles, the catalysts were analyzed by XRD, and the results are shown in Figure S1. Compared with the fresh Cu/CuO-BC patterns, the diffraction peaks of the Cu and CuO loaded on BC decreased slightly after usage, which might be attributed to the leaching of copper ions. The positions of diffraction peaks of Cu/CuO-BC remained unchanged before and after PMS activation, indicating that the catalyst was stable during repeating utilization.

The Recyclability of Cu/CuO-BC
The reusability of the catalyst is closely related to practical applicability. The recyclability of Cu/CuO-BC was evaluated by a recycling experiment. As shown in Figure 5, the TC degradation efficiency was reduced from 93.2% to 83.5% after using it five times. The results indicated that the Cu/CuO-BC prepared in this study exhibited relatively good reusability. In addition, the leaching amount of metal ions in the solution was detected by ICP-MS. The results showed that the leaching amount of Cu 2+ was 0.22 mg/L, much lower than the Chinese Sanitary Standard for Domestic Drinking Water GB 5749-2022 [36]. After five consecutive cycles, the catalysts were analyzed by XRD, and the results are shown in Figure S1. Compared with the fresh Cu/CuO-BC patterns, the diffraction peaks of the Cu and CuO loaded on BC decreased slightly after usage, which might be attributed to the leaching of copper ions. The positions of diffraction peaks of Cu/CuO-BC remained unchanged before and after PMS activation, indicating that the catalyst was stable during repeating utilization.

The Recyclability of Cu/CuO-BC
The reusability of the catalyst is closely related to practical applicability. The bility of Cu/CuO-BC was evaluated by a recycling experiment. As shown in Figu TC degradation efficiency was reduced from 93.2% to 83.5% after using it five tim results indicated that the Cu/CuO-BC prepared in this study exhibited relative reusability. In addition, the leaching amount of metal ions in the solution was dete ICP-MS. The results showed that the leaching amount of Cu 2+ was 0.22 mg/L, muc than the Chinese Sanitary Standard for Domestic Drinking Water GB 5749-2022 [3 five consecutive cycles, the catalysts were analyzed by XRD, and the results are sh Figure S1. Compared with the fresh Cu/CuO-BC patterns, the diffraction peaks o and CuO loaded on BC decreased slightly after usage, which might be attribute leaching of copper ions. The positions of diffraction peaks of Cu/CuO-BC remai changed before and after PMS activation, indicating that the catalyst was stable repeating utilization.

Contribution of Active Species for TC Degradation
In order to investigate the contribution of active species to the TC degradation by PMS activated with Cu/CuO-BC, typical chemical probes (trapping agents) were used to identify the active species during the reaction process. Ethanol (EtOH) was used as a trapping agent for SO4 •− (kSO4•− = (1.6-7.8) × 10 7 /M·s) and ·OH(k·OH = (1.2-2.8) × 10 9 /M·s), the tertiary butanol (TBA) was employed as a trapping agent for ·OH (k·OH = (3.8-7.6) × 10 8 /M·s), and furfuryl alcohol (FFA) was selected as trapping agents for 1 O2 (k1O2 = 1.2 × 10 8 /M·s) and ·OH (k·OH = 1.5 × 10 10 /M·s) [39]. Additionally, benzoquinone (BQ) was used for trapping O2 •− (kO2·− = 9.8 × 10 8 /M·s) and ·OH (k·OH = 6.6 × 10 9 /M·s) [40]. The results are shown in Figure 7. After adding 500 mM EtOH and TBA to the reaction solution, the TC removal efficiency decreased from 93.2% to 28.2% and 45.4% within 30 min, respectively. It indicated the TC degradation was inhibited significantly, indicating the SO4 •− and ·OH played an important role in the TC degradation. Furthermore, the addition of FFA and BQ resulted in a decrease in TC removal efficiency-as low as 24.8% and 20.9%, respectively. The effect of Cl − on TC degradation was small. As the Cl − concentration was increased from 0 to 2.0 mM, the degradation efficiency of TC decreased from 93.2% to 86.7% at 30 min. The degradation efficiency decreased by only 6.2%. In this process, Cl − would react with the active species to generate Cl-containing oxidants. The co-existing Cl − in solution might converted to Cl 2 •− , Cl • , and HClO (Equations (1)-(5)). The generated Cl 2 •− (E 0 = 2.0 V) and Cl • (E 0 = 2.4 V) possessed slightly weaker oxidation capacity than SO 4 •− (E 0 = (2.5-3.1 V)) [37] but were still capable of degrading TC. Thus, the system was less effective in degrading TC but still maintained a high removal rate. However, HCO 3 − in solution was able to inhibit the degradation of TC significantly. The degradation efficiency of TC decreased from 93.2% to about 50% when the concentration was increased from 0 to 2.0 mM. The HCO 3 − not only consumed the SO 4 •− in solution but also generated CO 3 with less oxidation ability (E 0 = 1.78 V) than the ·OH (Equations (6) and (7)) [38]. The concentration of the strongly oxidative species in the system is reduced.

•−
As shown in Figure 6c, (8) and (9)), which in turn reduced the catalytic activity. In the present experiment, the degradation of TC was not significantly inhibited, probably because H 2 PO 4 − reduced the activation energy of PMS, accelerated the production of active species, and slowed down the inhibitory effect of H 2 PO 4 − on TC degradation. In addition, the degradation efficiencies of TC within 30 min were 93.2%, 77.2%, 65.2% and 59.5%, respectively, at HA concentrations of 0, 2.0, 4.0 and 8.0 mg/L. As the concentration of HA increased, the inhibition effect was enhanced. The reactive functional groups, such as amine, carboxyl and phenolic hydroxyl groups in HA, reacted with the active species [39]. As a result, the steady-state concentration of the active species was reduced, and the degradation efficiency of TC was inhibited.

Contribution of Active Species for TC Degradation
In order to investigate the contribution of active species to the TC degradation by PMS activated with Cu/CuO-BC, typical chemical probes (trapping agents) were used to identify the active species during the reaction process. Ethanol (EtOH) was used as a trapping agent for SO 4 •− (k SO4•− = (1.6-7.8) × 10 7 /M·s) and ·OH(k ·OH = (1.2-2.8) × 10 9 /M·s), the tertiary butanol (TBA) was employed as a trapping agent for ·OH (k ·OH = (3.8-7.6) × 10 8 /M·s), and furfuryl alcohol (FFA) was selected as trapping agents for 1 O 2 (k 1O2 = 1.2 × 10 8 /M·s) and ·OH (k ·OH = 1.5 × 10 10 /M·s) [39]. Additionally, benzoquinone (BQ) was used for trapping O 2 •− (k O2·− = 9.8 × 10 8 /M·s) and ·OH (k ·OH = 6.6 × 10 9 /M·s) [40]. The results are shown in Figure 7. After adding 500 mM EtOH and TBA to the reaction solution, the TC removal efficiency decreased from 93.2% to 28.2% and 45.4% within 30 min, respectively. It indicated the TC degradation was inhibited significantly, indicating the SO 4 •− and ·OH played an important role in the TC degradation. Furthermore, the addition of FFA and BQ resulted in a decrease in TC removal efficiency-as low as 24.8% and 20.9%, respectively. Considering the trapped ·OH by FFA and BQ, the O 2 •− and 1 O 2 played a minor role in TC degradation.

The Possible Generation Mechanism of Active Species
In order to investigate the possible generation mechanism of active species, X-ray photoelectron spectroscopy (XPS) analysis was carried out on the virginal and used Cu/CuO-BC. The full spectrum of Cu/CuO-BC is shown in Figure S2. The main elements of Cu/CuO-BC were Cu, O, N and C, which remained unchanged before and after PMS activation, indicating the composition stability of Cu/CuO-BC. High-resolution XPS spectra of the metal element Cu of the virginal and used Cu/CuO-BC are shown in Figure 8. The Cu 2p spectrum of virginal Cu/CuO-BC catalyst has a characteristic peak at 932.6 and 934.2 eV, which represents the characteristic peaks of Cu 0 and Cu 2+ [41]. The composing proportion of Cu 0 (43.9%) was approximately twice that of Cu 2+ (22.7%), revealing the main existing species of Cu 0 in the catalyst, which was consistent with the XRD phase analysis. In the Cu 2p spectrum of the used Cu/CuO-BC, the characteristic peak at 932.6 eV was ascribed to Cu 0 , while the characteristic peaks at 933.3 eV and 934.2 eV were corresponded to Cu + and Cu 2+ , respectively. Notably, the proportion of Cu 2+ increased to 37.15% after the reaction, indicating the conversion of Cu 2+ during the TC degradation process. In addition, the characteristic peak of Cu + was detected, suggesting the occurrence of a reduction reaction from Cu 2+ . Meanwhile, Cu + was also oxidized to Cu 2+ by PMS, thus maintaining high catalyst activity via the valence cycle of Cu + and Cu 2+ .

The Possible Generation Mechanism of Active Species
In order to investigate the possible generation mechanism of active species, X-ray photoelectron spectroscopy (XPS) analysis was carried out on the virginal and used Cu/CuO-BC. The full spectrum of Cu/CuO-BC is shown in Figure S2. The main elements of Cu/CuO-BC were Cu, O, N and C, which remained unchanged before and after PMS activation, indicating the composition stability of Cu/CuO-BC. High-resolution XPS spectra of the metal element Cu of the virginal and used Cu/CuO-BC are shown in Figure 8. The Cu 2p spectrum of virginal Cu/CuO-BC catalyst has a characteristic peak at 932.6 and 934.2 eV, which represents the characteristic peaks of Cu 0 and Cu 2+ [41]. The composing proportion of Cu 0 (43.9%) was approximately twice that of Cu 2+ (22.7%), revealing the main existing species of Cu 0 in the catalyst, which was consistent with the XRD phase analysis. In the Cu 2p spectrum of the used Cu/CuO-BC, the characteristic peak at 932.6 eV was ascribed to Cu 0 , while the characteristic peaks at 933.3 eV and 934.2 eV were corresponded to Cu + and Cu 2+ , respectively. Notably, the proportion of Cu 2+ increased to 37.15% after the reaction, indicating the conversion of Cu 2+ during the TC degradation process. In addition, the characteristic peak of Cu + was detected, suggesting the occurrence of a reduction reaction from Cu 2+ . Meanwhile, Cu + was also oxidized to Cu 2+ by PMS, thus maintaining high catalyst activity via the valence cycle of Cu + and Cu 2+ .

The Possible Generation Mechanism of Active Species
In order to investigate the possible generation mechanism of active species, X-ray photoelectron spectroscopy (XPS) analysis was carried out on the virginal and used Cu/CuO-BC. The full spectrum of Cu/CuO-BC is shown in Figure S2. The main elements of Cu/CuO-BC were Cu, O, N and C, which remained unchanged before and after PMS activation, indicating the composition stability of Cu/CuO-BC. High-resolution XPS spec tra of the metal element Cu of the virginal and used Cu/CuO-BC are shown in Figure 8 The Cu 2p spectrum of virginal Cu/CuO-BC catalyst has a characteristic peak at 932.6 and 934.2 eV, which represents the characteristic peaks of Cu 0 and Cu 2+ [41]. The composing proportion of Cu 0 (43.9%) was approximately twice that of Cu 2+ (22.7%), revealing the main existing species of Cu 0 in the catalyst, which was consistent with the XRD phase analysis. In the Cu 2p spectrum of the used Cu/CuO-BC, the characteristic peak at 932.6 eV was ascribed to Cu 0 , while the characteristic peaks at 933.3 eV and 934.2 eV were cor responded to Cu + and Cu 2+ , respectively. Notably, the proportion of Cu 2+ increased to 37.15% after the reaction, indicating the conversion of Cu 2+ during the TC degradation process. In addition, the characteristic peak of Cu + was detected, suggesting the occur rence of a reduction reaction from Cu 2+ . Meanwhile, Cu + was also oxidized to Cu 2+ by PMS thus maintaining high catalyst activity via the valence cycle of Cu + and Cu 2+ .   Based on the above results, the generation mechanism of active species via PMS activation with Cu/CuO-BC was proposed, as shown in Figure 9. The Cu 0 site in Cu/CuO-BC coordinated with TC through the surface Cu-N bond, where the N site belonged to the amino group (-NH 2 ) in TC [42]. PMS could directly interact with the newly generated complexes with electron transfer from PMS to Cu 0 , thus facilitating PMS activation. Meanwhile, the Cu 0 converted to Cu + via redox reaction, producing ·OH and SO 4 •− (Equations (10) and (11)). Also, Cu + could directly react with HSO 5 − to generate SO 4 •− and ·OH (Equation (12)). Afterwards, the SO 4 •− could further combine with H 2 O to generate ·OH (Equation (11)). In addition, BC-based catalysts can produce O 2 •− and 1 O 2 via Equations (13) and (14). Then, the formed H 2 O 2 reacted with ·OH to form HO 2 · (Equation (15)), which in turn converted to O 2 •− (Equation (16)) [43]. Importantly, the Cu-N bridge formed between Cu 2+ and TC, which accelerated PMS activation and electrons transferred to the Cu 2+ sites, reducing Cu 2+ to Cu + (Equation (17)). Consequently, the Cu + /Cu 2+ cycle on the catalyst surface maintained the high utilization of PMS, contributing to the effective degradation of TC.

SO · + H O → H + SO +· OH (11)
Cu + HSO → Cu + SO · + OH (12) • Cu + HSO → Cu + SO · + H (17) Figure 9. Mechanism of the process of activation of PMS by Cu/CuO-BC to produce active species. Figure 9. Mechanism of the process of activation of PMS by Cu/CuO-BC to produce active species. ·OH

Chemicals
The peanut shell used in this study is the peanut shell waste of agricultural products. The solutions were prepared with ultrapure water. Copper sulfate (CuSO 4 ·5H 2 O) and TC with analytical grade were purchased from Shanghai Zehuasheng Technology Co., LTD (Shanghai, China) and Tianjin Biaotai Technology Development Co., LTD (Tianjin, China), respectively. Ethanol and PMS were supplied by Tianjin Jinke Fine Chemical Research Institute. Tert-butyl alcohol (TBA) and sodium dihydrogen phosphate (NaH 2 PO 4 ) were also analytically pure, which were obtained from Tianjin Damao chemical reagent factory. Sodium bicarbonate (NaHCO 3 ) and sodium chloride (NaCl) were purchased from Tianjin Biaotai and Guangfu Technology Development Co., LTD (Tianjin, China). Furfuryl alcohol and benzoquinone were purchased from Aladdin Reagent (Shanghai) Co., LTD (Shanghai, China).

Preparation of Cu/CuO-BC Catalysts
In the first stage, peanut shell was used to prepare biochar-based catalyst. The peanut shells were weighed and washed with deionized water 3 times and placed in an oven at 60 • C to dry to a constant weight. Then, peanut shells were crushed by a ball mill to obtain a finer peanut shell powder, which was sieved through 100 mesh and then placed in a tube furnace for pyrolysis with a heating rate of 10 • C/min, pyrolysis temperature of 300 • C and pyrolysis time of 2 h. The black powder was removed after the tube furnace cooled down naturally to room temperature. In the second stage, modified BC was prepared by impregnation pyrolysis. After the beaker was filled with 1.0 g of peanut-shell-derived powder, 200 mL of CuSO 4 solution with 0.1575 M was added to make the mass ratio (2) of Cu to the powder. The mixture was stirred continuously with a magnetic stirrer for 24 h. The supernatant was removed after standing, and the impregnated powder was dried and placed in a tube furnace for pyrolysis at 500, 600 and 700 • C with a heating rate of 10 • C/min and a pyrolysis time of 2 h. The preparation was completed when the tube furnace cooled down naturally to room temperature. For better comparison with the modified biochar, the powder obtained from the first stage also experienced re-pyrolysis at 500 • C without impregnation, which was named BC. The acquired modified biochar-based catalysts were marked as Cu/CuO-BC500, Cu/CuO-BC600 and Cu/CuO-BC700, respectively.

Characterization
The prepared BC and Cu/CuO-BC were analyzed by scanning electron microscopy (TESCAN MIRA LMS, Brno, Czech Republic) for microscopic morphology. The samples were scanned under energy scattering spectrometer (EDS) to detect the elemental composition on BC and Cu/CuO-BC surface. The specific surface area and pore size distribution of the BC-based catalysts were determined using a specific surface analyzer (McMurray-Tick Gemini VII 2390, McMurray, PA, USA). The surface functional groups of catalysts were analyzed using Fourier infrared spectroscopy (Thermo Fisher Nicolet iS10, Waltham, MA, USA). X-ray diffraction (Rigaku Smartlab SE, Akishima, Japan) was used to determine the structure and crystalline phase of the catalysts with a conventional angle of 10-80 • and a scanning speed of 10 • /min. X-ray photoelectron spectroscopy (Thermo Scientific ESCALAB 250Xi, Waltham, MA, USA) was used to detect the elemental composition of the catalysts and their valence structure. Electron paramagnetic resonance (EPR) spectroscopy measurement was conducted by a spectrometer (Bruker E500, Mannheim, Germany) at room temperature in the PMS system mediated by Cu/CuO-BC.

TC Degradation Experiment
In a 250 mL conical flask, 10 mL of 200 mg/L TC solution was added, diluted to 100 mL with deionized water, shaken well and then 0.01 g of catalyst powder was added. The catalyst was ultrasonically dispersed using an ultrasonic cleaner and then sampled at 300 r/min at a reaction temperature of 25 • C after stirring for 30 min to reach adsorption equilibrium.
PMS solution (2.0 mL) with a concentration of 25 mM was added after adsorption equilibrium, and samples were taken at 5, 10, 15, 20 and 30 min, respectively. The samples were filtered through a 0.22 µm polyethersulfone membrane, and the reaction was terminated by adding the same volume of 0.1 M sodium thiosulfate and then loaded into liquid phase vials for measurement. The TC concentration was detected by ultra-performance liquid chromatography (HPLC, SIL-30AC) with a C18 column (15 cm × 4.6 mm) following mobile phases: A: 0.1% formic acid aqueous solution; B: acetonitrile (chromatographic purity), VA:VB = 85:15, with gradient elution; flow rate: 0.1 mL/min; column temperature: 30 • C; injection volume: 10. The effects of co-existing anions on TC degradation were studied in the presence of single NaCl, NaHCO 3 and NaH 2 PO 4 . The effect of NOM on TC removal was investigated with humic acid. The TC relative concentration was expressed as C t /C a or C t /C 0 , where the C t , C a and C 0 were the concentrations of adding PMS for t min, reaching adsorption equilibrium and initial concentration, respectively. The pseudo-first-order kinetic model was described as ln(C t /C a ) = −K × t, where K is the rate constant and t is the degradation time.

Conclusions
In this study, Cu/CuO-BC catalysts were optimally prepared and characterized by a modified impregnation pyrolysis method with peanut shells as raw material. The research focused on the performance and mechanism of Cu/CuO-BC for TC degradation by PMS activation. The successful loading of Cu/CuO and the increased specific surface area (192.64 m 2 /g) facilitated the TC removal, achieving the degradation efficiency of 93.2% at 30 min with the kinetic constant of 0.0921 min −1 . The inhibition of TC degradation by co-existing anions (0.5-2.0 mM) followed the order of HCO 3 − > H 2 PO 4 − > Cl − . Also, HA had an inhibitory effect on TC degradation. The free radical (·OH and SO 4 •− ) pathway played a dominant role in TC degradation. The ·O 2 − and 1 O 2 were also involved in TC removal. The cycling of Cu + /Cu 2+ on the catalyst surface kept the high activity and good recyclability. The TC degradation efficiency was reduced from 93.2% to 83.5% after using five times. The prepared Cu/CuO-BC in this study was a highly active and environmentally friendly catalyst, which was of great value for new compound degradation in water.