Optimized Cu₂O-{100} facet for generation of different reactive oxidative species via peroxymonosulfate activation at specific pH values to efficient acetaminophen removal

1 Introduction

Recently, pharmaceutical and personal care products (PPCPs) detected in water bodies are posing serious threat to people’s health. As one of the typical PPCPs, acetaminophen (ACE) has been heavily applied to treat fever and relieve pain, which can be discharged into the surroundings by various paths and will endanger water circumstances and human bodies [1]. However, the removal of ACE from waste water is difficult to achieve through traditional sewage treatment processes due to its trace-level concentration and complex structure. Accordingly, it is quite essential to seek for advanced degradation systems and techniques to remove ACE from water.

Sulfate radical-based advanced oxidation processes (SR-AOPs) have been considered as the most promising processes to effectively remove refractory organic pollutants from wastewater. Various radicals and non-radials, such as sulfate radicals (SO₄˙⁻), hydroxyl radicals (˙OH), and singlet oxygen (¹O₂), are generated by activation of persulfate (PS, S 2 O 8 2 − ) or peroxymonosulfate (PMS, HSO 5 − ) during SR-AOPs, which mostly possess strong oxidized ability to achieve complete degradation of pollutants [2,3]. In general, PS and PMS can be activated with diversified degradation mechanisms via metal-based catalysts, thermal treatment, ultraviolet/alkali activation, and supersonic wave, which make it more difficult to control [3,4]. Meanwhile, water quality parameters (such as solution pH, impurity ions, and natural organic matter [NOM]) also influence the activation of PS/PMS, thereby, influencing pollutant degradation during water treatment. For instance, the formation and species of reactive oxidative species (ROS) are greatly dependent on solution pH [5]. Additionally, SO₄˙⁻ possesses a higher oxidation potential in contrast to conventional ˙OH under neutral pH, while SO₄˙⁻ and ˙OH display similar oxidation abilities in acidic condition [6]. Most of all, the reasonable design of metal-based catalysts, as the key factor for PMS activation, can optimize the degradation process with improved degradation rate [5]. Over the past decade, many strategies such as morphology/structure control, facet engineering, fabricating heterojunction, and element doping have been employed to achieve highly efficient catalysis [2,7]. For instance, the catalytic performances of Co₃O₄ catalysts in heterogeneous system for PMS activation depend on their morphologies and structures (e.g., nanospheres, nanotubes, nanorods, nanosheets, nanorings, etc.) [8]. Specifically, the crystal facets regulation is to change the exposed facets of catalysts with different geometric and electron structure. Precise regulation of crystal facets greatly improves catalytic activities and electron structure properties of catalysts [9]. On one hand, different crystal facets have different electron transfer rates and thus influence oxidation/reduction potentials. On the other hand, different crystal facets with different atomic arrangement and dangling bonds can alter the adsorption energy of target organic compounds on catalyst, consequently affecting its catalytic activities. For instance, Wu et al. reported that preferential growth of the cobalt (200) in Co@N–C exhibited excellent catalytic activities for PMS activation on the removal of bisphenol A [7]. Similarly, Hao et al. demonstrated that Fe(ii) on the {001} facets of hematite possessed a higher reactivity and electron transport efficiency than that on the {110} facets [9].

Of the various catalysts available to date, cuprous oxide (Cu₂O) is easy to realize exposed facet regulation, which also owns inherent advantages including low cost, excellent electron transport properties, and unique structures [10]. Most importantly, it has been reported that Cu₂O is a typical non-radical mechanism catalyst to efficiently activate PMS for organic pollutants degradation. Therefore, it is very important to develop tunable crystal facets of Cu₂O to realize highly efficient catalysis, clarify the effect of different exposed facets on catalysis, and unveil the activation mechanism under different water qualities, but the related research is yet ambiguous. Thus, it is essential to design Cu₂O catalysts with superior exposed specific facets for SR-AOPs in wastewater treatment and disclose the mechanism of ROS generation induced by Cu species for better catalytic activity.

Herein Cu₂O catalysts with different exposed facets are successfully synthesized, and certain characterizations have been conducted to verify their microscopic structure, which exhibits excellent degradation performances for ACE removal via PMS activation. The specific objectives are to: (i) develop Cu₂O catalysts with different exposed facets for ACE degradation via PMS activation, and investigate the influences of solution pH on ACE removal, (ii) clarify the structure–activity relationship of Cu₂O with different exposed facets and HSO 5 − by density functional theory (DFT) calculations, (iii) identify the main ROSs during reactions under different solution pH via quenching tests and electron paramagnetic resonance (EPR) analysis, (iv) explore the changes of Cu species affected by various pH, and (v) elucidate the mechanism for enhanced organic pollutants degradation by PMS activation. This work can deepen the understanding of PMS activation induced by Cu-based catalysts and will provide a new direction for non-radical processes in wastewater treatment.

2 Experimental method

3 Results and discussion

Figure 1(a1–c1) and (a2–c2) displayed the ESEM images of Cu₂O catalysts with typical cubic, vertex-truncated octahedral, and octahedral structures, which were obtained with the amount of PVP increased. Obviously, all these catalysts owned the quite uniform size of about 700 nm, and their corresponding three-dimensional models are provided in the insets of Figure 1(a2–c2). The exposed surfaces of the cubes, vertex-truncated octahedrons, and octahedrons were enclosed of six {100} facets, six {100} and eight {111} facets, and eight {111} facets (Figure 1(a2–c2)), which could be verified by the lattice fringes in the HRTEM images (Figure 1(a3–c3)).

Figure 1

Morphology characterization of Cu₂O catalysts. The SEM images of the (a₁ and a₂) cubic Cu₂O, (b₁ and b₂) vertex-truncated Cu₂O, and (c₁ and c₂) octahedral Cu₂O, and insets were the corresponding three-dimensional models. HRTEM images of (a₃) cubic Cu₂O, (b₃) vertex-truncated octahedron Cu₂O, and (c₃) octahedral Cu₂O.

To study the structural information and surface chemical state, XRD and XPS spectra of as-obtained catalysts were measured. As displayed in Figure 2, several apparent diffraction peaks at 29.5°, 36.4°, 42.3°, and 61.3° were observed for all samples, corresponding to the (110), (111), (200), and (220) planes of cuprite Cu₂O (JCPDS No. 77-0199) [10]. Moreover, the XPS survey spectra of different samples in Figure S1a confirmed the existence of Cu and O in all the three Cu₂O catalysts. By comparison, more active Cu atoms were exposed on the surface of cubic Cu₂O with a Cu/O atomic ratio of 0.99, which was higher than that of vertex-truncated Cu₂O (0.66) and octahedral Cu₂O (0.58). This probably resulted from the difference in their exposed crystal facets, which then induced the different PMS activation performances during reactions [8]. In Figure S1b, two peaks were observed for the high-resolution spectra of O 1s at 530.5 and 531.4 eV, assigned to the lattice oxygen and surface adsorbed H₂O, respectively [16,17]. The core-level Cu 2p spectrum in Figure S1c exhibited two dominant peaks centering at 932.7 and 952.4 eV, which belonged to Cu 2p_3/2 and Cu 2p_1/2 of Cu⁺ [18]. Besides, two weak peaks at 934.7 and 955.2 eV were assigned to Cu²⁺, which might be resulted from slight oxidation of Cu₂O [16].

Figure 2

XRD patterns of cubic, vertex-truncated, and octahedral Cu₂O catalysts.

The degradation of ACE was investigated through PMS activation over different catalysts. As displayed in Figure 3a, the concentration of ACE was not decreased after adsorption for 10 min, which suggested that the cubic Cu₂O had no capacity to adsorb ACE. Meanwhile, the control experiment with only PMS addition indicated that PMS could not degrade ACE directly. However, an obvious removal of ACE was observed after adding cubic Cu₂O catalyst with admirable efficiency of 100% in 7 min. In contrast, octahedral Cu₂O and vertex-truncated octahedral Cu₂O exhibited inferior degradation efficiency of 88.23 and 94.55% in 10 min, respectively. Pseudo first-order kinetic rate constants (k ₁) of ACE degradation over various catalysts were evaluated and demonstrated in Figure 3b. The relevant k ₁ of cubic Cu₂O was 0.61 min⁻¹, which was 1.74 and 2.65 times of vertex-truncated octahedral (0.35 min⁻¹) and octahedral Cu₂O (0.23 min⁻¹), verifying its higher kinetic reaction rate. These results illustrated that cubic Cu₂O possessed superior ability to activate PMS and thus improved the oxidation degradation of refractory organic contaminants.

Figure 3

(a) Different catalyst-PMS systems for ACE removal and (b) the corresponding pseudo first-order kinetic model fitting. (c) The influence of different pH on ACE removal over cubic CuO₂ by PMS activation and the inset was the corresponding degradation efficiency with 10 min. Influences of (d) Cl⁻, (e) coexisting cations, and (f) HA on ACE removal in various cubic CuO₂ systems. Reaction conditions: [ACE]₀ = 10 μmol/L, [PMS]₀ = 100 μmol/L, [catalyst]₀ = 0.1 g/L, [Cl⁻] = 1–10 mmol/L, [cation] = 2 mmol/L, [HA] = 0.25–1.0 mg/L, and solution pH 5.0 ± 0.2.

Since the generation of ROS was closely depended on the reaction system pH, we evaluated the effects of different initial solution pH on ACE degradation using cubic Cu₂O as targeted catalyst. As shown in Figure 3c, cubic Cu₂O exhibited the satisfactory degradation efficiency for ACE over a wide pH range from 3 to 11. Specially, ACE could be almost completely eliminated from the solution at pH 3, 5, and 11. Thus, the desired cubic Cu₂O could work as a Janus catalyst in both acid and alkaline solvent environment.

Subsequently, the influences of co-existing ions and NOM on ACE degradation were also explored (Figure 3d–f). It could be noted that the degradation rate of ACE gradually increased with the Cl⁻ concentrations increasing from 1 to 10 mmol/L (Figure 3d). This was because the Cl⁻ anions might react with HSO 5 − /SO₄˙⁻/˙OH to yield active chlorine species, such as HClO, ClO˙, and Cl˙, which possessed high selectivity and preferred to react with electron-rich moieties containing anilines, phenols, and olefins, thereby resulting in strong oxidation capacity for ACE removal [19]. As observed from Figure 3e, Na⁺ and Mg²⁺ ions showed no obvious effect during degradation reactions. However, the ACE degradation rate was inhibited by Ca²⁺ ions, which resulted due its electrical double layer compression effect and ion exchange capacity [20]. Figure 3f displayed that the ACE degradation efficiency decreased as the humic acid (HA) concentration increased from 0.25 to 1 mg/L, which was due to the fact that HA was a quencher of free radicals. Thus, cubic Cu₂O with exposed {100} facets exhibited preferable degradation efficiency and significant adaptation in wide pH range as well as high salt condition.

To unveil the high activity from cubic Cu₂O and clarify the roles of different exposed facets on PMS activation, DFT calculation was performed. As shown in the insets of Figure 4a and b, structural optimizations of bulk Cu₂O were conducted, and two representative models with {100} and {111} exposed facets were obtained, which were denoted as Cu₂O-{100} (cubic Cu₂O) and Cu₂O-{111} (octahedral Cu₂O). Correspondingly, the electronic densities of state (DOS) for the two different models were also displayed in Figure 4a and b. Interestingly, the charge density state of Cu₂O-{100} was across the Fermi level, suggesting its metallic property and superior electron transfer efficiency. While, the DOS spectra of Cu₂O-{111} belonged to the characteristic feature of semiconductor, thus resulting in the decrease in conductive carriers. That is, rational morphology control effectively regulated the electron migration in cubic Cu₂O with {100} exposed facets, which promoted the conductivity performance and interfacial reactions. The electron transfer property was further proved by electrochemical impedance spectroscopy in Figure S2, of which the semicircle diameter for cubic Cu₂O was smaller than that of octahedral Cu₂O and vertex-truncated Cu₂O, indicating that the {100} facet for cubic Cu₂O benefited for lowering the electron transfer resistance. Besides, the d-band center of Cu₂O-{100} was calculated to be −2.27 eV, while it shifted to −2.37 eV for Cu₂O-{111} after forming octahedral structure, hinting that the stronger adsorption ability of Cu₂O-{100} for HSO 5 − than that of Cu₂O-{111}.

Figure 4

The DOS and atomic structure models of (a) Cu₂O-{100} and (b) Cu₂O-{111}. The DOS of (c) Cu₂O-{100} and (d) Cu₂O-{111} after adsorbing HSO 5 − molecules. The adsorption energy and CDD for HSO 5 − on (e) Cu₂O-{100} and (f) Cu₂O-{111}, respectively. The dark yellow and cyan regions represented the electron accumulation and charge depletion, respectively.

Furthermore, DFT calculations were employed to investigate the interaction between HSO 5 − and Cu₂O catalysts with different facet surfaces. It could be observed from Figure 4c that, after adsorbing HSO 5 − molecules, no obvious change across the Fermi level was found for Cu₂O-{100}, demonstrating that the extraordinary electrical conductivity of cubic Cu₂O still remained. This meant that HSO 5 − molecules as electron acceptors could effectively capture electrons from cubic Cu₂O and be activated to produce sufficient radicals for pollutant degradation. With regard to Cu₂O-{111}/ HSO 5 − in Figure 4d, the obvious hybridization between O-p and Cu-4d orbitals occurred at the Fermi energy, which was beneficial for accelerating electron transport during the HSO 5 − adsorption process. Moreover, charge density difference (CDD) and the Bader charge analysis were carried out to deeply elucidate the surface electron transfer mechanism. After adding HSO 5 − molecules, the adsorption geometry structures of Cu₂O-{100}/ HSO 5 − and Cu₂O-{111}/ HSO 5 − were optimized and presented in Figure 4e and f. The adsorption energy of Cu₂O-{100} for HSO 5 − (−4.21 eV) was much higher than that of Cu₂O-{111} (−3.06 eV), suggesting the superior adsorption ability of Cu₂O-{100} toward HSO 5 − , which was consistent with the d-band center results. Note that the charge depletion (the cyan region) occurred between the Cu–O and S–O bonds, while the electron accumulation (the dark yellow region) emerged around the bottom three O atoms in HSO 5 − , resulting in an efficient activation of PMS on the surface of Cu₂O. According to the Bader topological calculation, it was found that HSO 5 − molecule captured 2.09 e⁻ and 1.99 e⁻ from the substrate Cu₂O-{100} and Cu₂O-{111}, respectively. This phenomenon indicated that valence electrons on cubic Cu₂O could be effectively separated and transferred to bond with HSO 5 − . Correspondingly, the O–O bond length of HSO 5 − on Cu₂O-{100} (1.473 Å) was slightly increased than that on Cu₂O-{111} (1.467 Å). These findings confirmed that HSO 5 − molecules were preferably activated by cubic Cu₂O to generate more ROS. Therefore, DFT calculation results proved that Cu₂O-{100} could provide more fluent electron transfer and contribute to PMS activation for improving the catalytic activity.

To unveil degradation mechanism of ACE in cubic Cu₂O/PMS system at different initial pH, quenching experiments were carried out to confirm the major ROS generated in suspension during ACE oxidation (Figure 5a–c). TBA, EtOH, and FFA were used to quench ˙OH, SO₄˙⁻/˙OH, and ¹O₂, respectively [21,22]. N₂ was bubbled into suspensions to eliminate the interference of dissolved O₂ on ACE degradation. With the addition of TBA, EtOH, or FFA into the reaction system at pH 3 (Figure 5a), respectively, the removal of ACE was completely inhibited, which indicated that ˙OH and ¹O₂ were majorly responsible for ACE degradation. A slight decrease in the removal efficiency for ACE was observed in cubic Cu₂O/PMS system after bubbling N₂, meaning that dissolved O₂ might not participate in oxidation reactions at pH 3. Under pH 5, the degradation efficiencies of ACE were 90.65, 90.63, and 0% with TBA, EtOH, and FFA as quenching agents (Figure 5b), respectively, which meant that the contribution of ¹O₂ and SO₄˙⁻/˙OH toward ACE removal were 90.63 and 9.37%. Figure 5b showed that ACE degradation was obviously suppressed in N₂ atmosphere with the removal efficiency of 26.95%. Thus, ¹O₂ played a vital role during ACE degradation at pH 5 and the action of dissolved O₂ was not neglected. In alkaline condition of pH 11, a similar result was obtained (Figure 5c), wherein the addition of TBA displayed no obvious impact on ACE removal, while the degradation efficiency decreased to 81.89% after introducing EtOH, further confirming that 18.11% of ACE was degraded by SO₄˙⁻ radicals. Once adding FFA into the catalytic system, the degradation of ACE was entirely suppressed, which demonstrated that ¹O₂ served as the major ROS during degradation at pH 11 and its contribution toward ACE removal was 81.89%. Additionally, 94% of ACE was eliminated at N₂ atmosphere, and the slight inhibition effect was similar to the results obtained from pH 3.

Figure 5

(a–c) Effects of different quenchers on ACE degradation in cubic Cu₂O/PMS system at (a) pH 3, (b) pH 5, and (c) pH 11. EPR spectra of DMPO-SO₄˙⁻, DMPO-˙OH, and TEMP-¹O₂ for PMS activation in N₂ atmosphere under (d) pH 3, (e) pH 5, and (f) pH 11.

To explore the generation of ROS for ACE degradation in cubic Cu₂O/PMS system, spin-trapping EPR spectra were performed at pH 5 in air atmosphere. As observed from Figure S3a, the typical four-line spectra with relative intensities of 1:2:2:1 were detected in different Cu₂O/PMS systems by using DMPO as trapping agent, demonstrating that DMPO-˙OH adducts were identified with hyperfine splitting constant of α _N = α _H = 14.9 G¹¹¹² [23]. And also, DMPO-SO₄˙^‒ adducts were found in all these systems, of which six special peaks with hyperfine splitting constants of α _N = 13.2 G, α _H = 9.6 G, α _H = 1.48 G, and α _H = 0.78 G could be identified to originate from SO₄˙^‒ radicals [23]. Moreover, three-line spectra of TEMP-¹O₂ adducts were found in different Cu₂O/PMS system (Figure S3b), verifying that ¹O₂ was produced and oxidized TEMP to TEMPO [24]. On comparing with vertex-truncated and octahedral Cu₂O catalysts, the EPR signals from DMPO-˙OH, DMPO-SO₄˙^‒, and TEMP-¹O₂ adducts were significantly enhanced in cubic Cu₂O/PMS system, implying that more ˙OH, SO₄˙^‒, and ¹O₂ were generated during PMS activation over cubic Cu₂O. Combining with the results from quenching experiments, ¹O₂ was proved as the dominant ROS for ACE oxidation in Cu₂O/PMS system. In order to confirm the generation pathways of ¹O₂, EPR techniques were also used to explore the changes in ROS in Cu₂O/PMS systems at different solution pH in N₂ atmosphere. Observing from Figure 5d, no EPR signals of DMPO-SO₄˙^‒ adducts were detected, while some ˙OH radicals were captured during ACE degradation at pH 3, which was probably due to Cu(iii) hydrolysis [25]. Interestingly, the elimination of dissolved O₂ prevented the generation of O₂˙^‒ in this system. And meanwhile, after adding TBA to cubic Cu₂O/PMS system to scavenge ˙OH radicals during PMS activation, no EPR signals of TEMP-¹O₂ adducts were detected (Figure S4). Thus, it could be speculated that the formation of ¹O₂ was derived from the disproportionation reaction of ˙OH similar to previous report [26]. Under pH 5, only ˙OH and SO₄˙^‒ were observed in the suspensions and are responsible for ACE oxidation (Figure 5e). By contrast, no TEMP-¹O₂ adducts emerged due to lack of O₂ source, which indicated the generation of ¹O₂ from the dissolved O₂. In alkaline condition of pH 11, all signals of TEMP-¹O₂, DMPO-SO₄˙^‒, and DMPO-˙OH were monitored (Figure 5f), confirming the generation of SO₄˙^‒, ˙OH, and ¹O₂ through direct activation of PMS by cubic Cu₂O. Besides, at pH 11, the ACE degradation showed no obvious change in N₂ atmosphere, which suggested that dissolved O₂ did not participate in reactions and the generation of ¹O₂ was derived from the SO₅˙^‒ hydrolysis or disproportionation reaction [27]. Overall, the desired cubic Cu₂O catalysts not only exhibited superior catalytic performances, but also adapted well to a wide pH range. Note that ¹O₂ was extremely important to organic pollutants removal in Cu₂O/PMS system, and the degradation mechanism verified that the non-radical process was vital for PMS activation by cubic Cu₂O.

The confocal Raman was further performed to observe in situ the interaction processes of the cubic Cu₂O/PMS system. As displayed in Figure 6a, three peaks at 880, 980, and 1,060 cm⁻¹ could be observed in the fresh PMS solution alone, which belonged to the stretching vibration of O–O bonds in HSO 5 − , and symmetrical stretching vibrations from SO₄ ²⁻ and HSO 5 − , respectively [28,29]. Besides, pure ACE exhibited a broad vibration band centering at 780 cm⁻¹, which was also observed in ACE/Cu₂O/PMS systems. Surprisingly, after the adsorption of ACE onto the metal sites of cubic Cu₂O, two new characteristic peaks appeared at approximately 558 and 1,092 cm⁻¹, which belonged to the Cu–N stretching vibrations and thus confirmed the formation of Cu–ACE complexes [28]. Similarly, the special peaks (Cu–N) also emerged in ACE/Cu₂O/PMS suspensions with different solution pH, meaning that ACE could interact well with cubic Cu₂O in a wide pH range, which benefited for improving the degradation activity.

Figure 6

(a) In situ Raman spectra of ACE, PMS, ACE/Cubic Cu₂O, and ACE/Cubic Cu₂O/PMS at different solution pH. The XPS spectra of (b) Cu 2p and (c) O 1s for cubic Cu₂O before and after degradation reactions. (d) The removal mechanisms in various pH for removing ACE.

Based on quenching experiments, EPR spectra, and in situ Raman spectra, the possible mechanisms for ACE degradation were proposed. Three alternative formation pathways of ROS under different pH were supposed. (i) For the degradation process in Cu₂O/PMS system at pH 3, Cu(i) would react with HSO 5 − to produce Cu(iii) (equation (4)), which was further hydrolyzed to acquire ˙OH and H⁺ ions (equation (5)). It has been reported that Cu(iii) could rapidly convert PMSO into PMSO₂ by the oxygen transfer reaction [30]. To validate the formation of Cu(iii) as intermediate products during degradation, PMSO was added into the cubic Cu₂O/PMS system. Consequently, the concentration of PMSO was reduced from 1.0 to 0.7 mmol/L, while an obvious increase in PMSO₂ concentration (from 0 to 0.3 mmol/L) was observed, which further proved the existence of Cu(iii) during PMS activation. Meanwhile, ¹O₂ was derived from the disproportionation reaction of ˙OH (equation (6)). At last, ¹O₂ and ˙OH worked as major active species to attack and destroy ACE molecules to other smaller oxidized products in acid solution. (ii) At pH 5, PMS mainly existed as HSO 5 − anions, which could be activated by Cu(i) to yield seldom ˙OH and SO₄˙⁻ radicals (equations (7) and (8)) [31]. While, SO₄˙⁻ and ˙OH radicals were not the main ROS for ACE removal as mentioned in Figure 5b. Based on the quenching experiments and EPR trapping results, we conjectured that Cu(i) could react with O₂ to form Cu(ii) and O₂˙⁻ (equation (9)), and then ¹O₂ originated from O₂˙^‒, which was produced by the self-reaction of in situ generated O₂˙^‒ (equations (10) and (11)) or reactions of O₂˙^‒ with ˙OH (equation (12)). (iii) Under pH 11, dissolved O₂ did not participate in the formation of ¹O₂ during reactions, and the dominant species for PMS in suspensions were SO₅ ²⁻ anions [32], which might react with HSO 5 − to form ¹O₂ (equation (13)). In addition, SO₄˙⁻ from Cu(i)-induced PMS activation would react with OH⁻ to generate ˙OH (equation (14)). Meanwhile, Cu(ii) could combine with HSO 5 − to generate SO₅˙⁻ (equation (15)), which promoted the generation of ¹O₂ (equations (16) and (17)). Finally, the electron transfer and chemical states of the fresh and used cubic Cu₂O after reacting with PMS activation were also investigated by XPS spectra, and the results are displayed in Figure 6b and c. It could be clearly noted that, by contrast to fresh cubic Cu₂O, the content of Cu(ii) in the copper species after oxidation degradation increased from 17 to 31%. This could be resulted from the generation process of ¹O₂, during which Cu(i) was oxidized to high valence copper species (e.g., Cu(ii) and Cu(iii)) after reacting with PMS, and simultaneously, Cu(ii) would also be reduced to Cu(i) through degradation reactions, forming a cyclic reaction [33]. Due to the poisoning of catalyst in actual reactions, Cu(ii) species could not be fully converted into Cu(i) after PMS activation. As for the O 1s spectra in Figure 6c, it could be noted that the ratio of adsorbed oxygen increased from 24% (fresh cubic Cu₂O) to 35% (used cubic Cu₂O). The phenomenon was probably due to the involvement of the OH⁻ in H₂O for PMS activation, resulting in the formation of more adsorbed oxygen species on the surface of catalysts. Thus, the XPS analysis provided powerful evidence with the possible mechanisms, which was definitely concluded in Figure 6d.

4 Conclusion

In summary, a potential cubic Cu₂O catalyst was successfully developed and employed to efficiently activate PMS for the refractory pollutants’ removal. The superior catalytic oxidation behaviors of cubic Cu₂O with {100} facets could be ascribed to its rapid electron transfer and strong PMS adsorption ability. Consequently, cubic Cu₂O displayed a higher degradation rate for ACE (0.61 min⁻¹) by PMS activation than that of vertex-truncated octahedral Cu₂O (0.35 min⁻¹) and octahedral Cu₂O (0.23 min⁻¹), and the highly-efficient catalytic activity could be achieved in a wide pH range of 3–11. The mechanistic study confirmed that ¹O₂ was dominant ROS during oxidation reactions, which was derived from PMS activation and mainly depended on the redox cycle of Cu species in cubic Cu₂O. This work shed new light on PMS-AOPs toward Cu-based catalysts, and offered a sustainable strategy for wastewater purification by non-radical reactions.