Degradation of Tetracycline Hydrochloride in Water by Copper–Iron Bioxide-Activated Persulfate System

Abstract

Advanced oxidation processes (AOPs) utilizing peroxymonosulfate (PMS) have emerged as a promising technology for organic pollutant degradation due to their distinct environmental advantages. In this study, copper–iron bimetallic oxide catalysts with varying ratios were synthesized via a co-precipitation method to activate PMS for degrading simulated tetracycline hydrochloride wastewater. The catalysts were characterized by scanning electron microscopy (SEM), X-ray diffraction (XRD), Fourier transform infrared spectroscopy (FTIR), and X-ray photoelectron spectroscopy (XPS). The effects of key parameters—including the PMS concentration, catalyst dosage, initial pH, and tetracycline hydrochloride concentration—on the degradation efficiency were systematically investigated. The results demonstrated that the CuFe(2)/PMS system exhibited the highest degradation efficiency. Under optimal conditions (20 mg/L tetracycline hydrochloride, 0.4 mM PMS, 0.5 g/L CuFe(2) catalyst, and pH 3), this system achieved a 94.12% degradation rate of tetracycline hydrochloride within 120 min. The electron paramagnetic resonance (EPR) tests and radical quenching experiments identified sulfate radicals (SO₄·⁻) as the predominant reactive species. Furthermore, the XPS analysis elucidated the persulfate activation mechanism, while the liquid chromatography–mass spectrometry (LC-MS) identified the potential degradation pathways and intermediate products of tetracycline hydrochloride.

1. Introduction

With the advancement of socioeconomic and technological development, the use of antibiotics in medical and aquaculture sectors has significantly increased, contributing notably to disease prevention and treatment [1]. Tetracycline antibiotics not only play a vital role in human health but are also widely used in animal feed to promote animal health and enhance growth efficiency [2]. In China, with the increase in their usage, tetracycline antibiotics enter the natural environment through various pathways, such as agriculture, medical applications, and animal farming. The monitoring data for China’s surface water environment show that the detection frequency and concentration levels of these antibiotics are relatively high. Concentrations of chlortetracycline and oxytetracycline in the Haihe River Basin reached 68.9 μg/L and 361 μg/L, respectively [3]. Among these, tetracycline hydrochloride (TCH) ranks as the second most widely used antibiotic globally [4], attributable to its straightforward synthesis, low cost, and excellent antibacterial efficacy [5]. However, TCH is not fully metabolized by animals and humans; a substantial portion is excreted into the environment as the unmetabolized parent compound through urine and feces [6,7]. Due to its low biodegradability and complex molecular structure [8], TCH residues in the environment are persistent, leading to bioaccumulation in the food chain and subsequent entry into the human body, posing significant health risks [9,10,11,12]. The current methods for TCH removal, including biodegradation [13], adsorption [14], and electrolysis [15], are hampered by their high costs and slow reaction rates [16,17]. Therefore, developing an efficient, cost-effective method for TCH degradation is of great practical importance.

In recent years, advanced oxidation processes (AOPs) utilizing persulfate have attracted significant interest for organic pollutant degradation owing to their high redox potential (2.6–3.1 V), prolonged radical lifespan (30–40 μs), and broad pH applicability (3–11) [18,19,20]. Persulfate activation (PS) generates highly reactive radicals, such as sulfate hydroxyl radicals (·OH), radicals (SO₄^·-), and superoxide radicals (O₂^·-) [21], which possess strong oxidizing capabilities, transforming pollutants into benign end products like CO₂ and H₂O [18] or non-toxic intermediates [20]. Compared to peroxydisulfate, peroxymonosulfate (PMS) is more readily activated, primarily due to its longer superoxide bond and asymmetric structure [22]. Various activation methods for PS, including ultrasound, UV light, electrochemistry, heat, and transition metals, have been extensively studied [23]. For instance, Li et al. [5] demonstrated effective TCH degradation using photoactivated PS, while Liu et al. [19] utilized electrochemical activation to achieve persulfate oxidation, effectively removing TCH from aqueous solutions. Gao et al. [24] showed that microwave energy can efficiently treat TCH by activating PS. Among these methods, transition metal activation of persulfate has drawn significant interest owing to its low cost and simplicity [18].

In homogeneous reactions, transition metal ions dissolve in a solution, making them difficult to recover and reuse. In contrast, heterogeneous transition metal catalysts exist in solid form, enabling easy solid–liquid separation and secondary recovery, which is highly beneficial for cost savings. Heterogeneous transition metal catalysts, such as metal–organic frameworks [25] and metal oxides [26], can effectively activate PMS, and have attracted extensive research attention. Compared with single-metal oxides, bimetallic oxides or mixed metal oxides may exhibit higher activity and stability in activating PMS due to the potential synergistic effects between different metals, while also suppressing metal leaching. Copper and iron oxides represent the common transition metal oxides in advanced oxidation technologies [27,28]. Wang et al. [29] synthesized copper–iron composites with varying Cu/Fe ratios via a hydrothermal method and employed them as catalysts to activate persulfate for ofloxacin removal, achieving a 96.2% removal rate within 30 min. Shi et al. [30] prepared a CuO@Fe₃O₄ magnetic material to activate PMS for degrading *p*-nitrophenol (PNP), achieving 96% degradation efficiency within 60 min. These results revealed significant bimetallic synergy between the copper and iron elements. Therefore, combining copper and iron to prepare mixed metal compounds represents an effective strategy for activating PMS.

Although numerous copper–iron bimetallic catalysts have been developed for PMS activation, traditional aqueous-phase-synthesized Fe/Cu bimetallic particles may cause substantial iron dissolution or oxidation [31]. Products obtained via the hydrothermal method exhibit high activity, excellent quality, and uniform distribution. However, this method has high requirements for the raw materials and incurs high manufacturing costs. While mechanical ball milling offers advantages, such as simple experimental equipment and easy control of synthesis elements, the products synthesized by this method often suffer from an uneven size distribution and irregular shapes. In contrast, the co-precipitation method involves introducing an appropriate precipitant into a solution containing two or more types of cations to form a precursor precipitate. After filtration, washing, drying, and calcination, the corresponding material is obtained. This method is characterized by simple equipment requirements and a straightforward operation, yielding products with high surface reactivity and good stability. However, in most studies, calcination is typically carried out in a muffle furnace without isolating air, which may lead to unintended oxidation of the catalyst.

Unlike previous methods, this study synthesized a copper–iron bimetallic catalyst via co-precipitation. Calcination was subsequently conducted in a tubular furnace under a continuous nitrogen flow, preventing atmospheric oxygen reactions during high-temperature treatment and avoiding unpredictable catalytic performance alterations. By pre-adjusting the copper-to-iron molar ratio, we tailored the Cu-Fe interactions. Characterization techniques (XRD, FTIR, XPS, and SEM) analyzed the catalyst’s structure, surface properties, and composition. We systematically examined the impacts of Cu-Fe ratio, pH, PMS dosage, tetracycline hydrochloride concentration, and catalyst dosage on the degradation performance. Electron paramagnetic resonance (EPR) identified the reaction-generated radicals, with quenching experiments determining the dominant species. An XPS analysis elucidated the metal valence state changes before and after the reactions, revealing the radical generation mechanism through PMS activation. To identify the degradation pathways and intermediate products of TCH, we used liquid chromatography–mass spectrometry (LC-MS). The analysis revealed degradation routes, such as demethylation, dehydroxylation, and ring-opening reactions. These insights contribute to the understanding of degradation mechanisms and support advancements in treating organic pollutants.

2. Materials and Methods

2.1. Chemicals and Reagents

All chemicals utilized in this study were of analytical grade and employed without further purification. Copper nitrate trihydrate (Cu(NO₃)₂·3H₂O), iron nitrate nonahydrate (Fe(NO₃)₃·9H₂O), potassium peroxymonosulfate (KHSO₅), and tert-butanol (C₄H₁₀O) were sourced from Shanghai Macklin Biochemical Co., Ltd. (Shanghai, China). Concentrated hydrochloric acid (HCl), sodium hydroxide (NaOH), sodium carbonate (Na₂CO₃), anhydrous ethanol (C₂H₆O), tetracycline hydrochloride (C₂₂H₂₅ClN₂O₈), and methanol (CH₃OH) were procured from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). All experimental solutions were prepared using ultrapure water.

2.2. Preparation of CuFe Catalysts

The copper–iron bimetallic oxide catalyst was synthesized via a co-precipitation method followed by calcination in a tube furnace under a nitrogen atmosphere. At room temperature, 0.9664 g of Cu(NO₃)₂·3H₂O and 6.464 g of Fe(NO₃)₃·9H₂O were dissolved in 100 mL of 0.5 mol/L Na₂CO₃ aqueous solution, resulting in the formation of a reddish-brown precipitate. This precipitate was then filtered and washed three to four times with deionized water and ethanol. After drying at 80 °C for 12 h, the material was calcined at 500 °C for 3 h in a nitrogen atmosphere, yielding a product with a Cu:Fe molar ratio of 1:4, designated as CuFe(1). By adjusting the molar ratios of Cu(NO₃)₂·3H₂O to Fe(NO₃)₃·9H₂O, copper–iron bimetallic oxides with Cu:Fe ratios of 1:2 and 4:1 were similarly prepared, designated as CuFe(2) and CuFe(3), respectively. The preparation flow chart is shown in Figure 1a.

2.3. Characterization Methods of CuFe Catalysts

To verify the successful synthesis of the catalyst, various characterization techniques were employed. Scanning electron microscopy–energy-dispersive X-ray spectroscopy (SEM-EDS, ZEISS Sigma500, ZEISS, Oberkochen, Germany) was utilized to analyze the surface morphology and elemental composition of the catalyst. Fourier transform infrared spectroscopy (FTIR, Nicolet, Thermo Fisher Scientific, Waltham, MA, USA) was applied to identify the functional groups present in the catalyst. The crystal structure was determined using an X-ray diffractometer (XRD, D8 ADVANCE, Bruker, Billerica, MA, USA). Changes in the chemical states of elements before and after the catalytic reaction were examined by X-ray photoelectron spectroscopy (XPS, Thermo Kalpha, Thermo Fisher Scientific, East Grinstead, UK). Electron paramagnetic resonance spectroscopy (EPR, EMXmicro, Bruker, Billerica, MA, USA) was used to detect the free radical signals. Additionally, a molecular analysis was performed using liquid chromatography–mass spectrometry (LC-MS, Thermo TSQ Quantis, Thermo Fisher Scientific, San Jose, CA, USA). Chromatographic separation was performed on a Thermo Scientific Accucore C18 column (150 mm × 2.1 mm, 2.6 μm) maintained at 40 °C. The mobile phase consisted of (A) 0.1% formic acid in water and (B) methanol at a flow rate of 0.3 mL/min with gradient elution as follows: 0–2 min, 10% B; 2–15 min, 10–95% B; 15–18 min, 95% B. The sample injection volume was 5 μL. The mass detection employed electrospray ionization (ESI) in the positive mode with a spray voltage of 3.5 kV. Full scans (m/z 100–600) were acquired using a TSQ Quantis mass spectrometer. Prior to analysis, the samples were filtered through 0.22 μm nylon membranes.

2.4. Degradation Experiments

Selection of the Optimal Catalyst: A PMS concentration of 0.1 mM was utilized alongside CuFe(1), CuFe(2), CuFe(3), and Cu(NO₃)₂·3H₂O concentrations of 0.5 g/L. The tetracycline hydrochloride solution concentration stood at 20 mg/L. The solution pH remained unadjusted, and samples were taken at 5 min intervals up to 120 min to gauge the tetracycline hydrochloride concentration, facilitating the determination of the most effective catalyst.

Exploring the Impact of pH on TCH Degradation Rate: Employing an optimal catalyst concentration of 0.5 g/L and a PMS concentration of 0.1 mM, the tetracycline hydrochloride solution concentration remained at 20 mg/L. pH variations were induced using different HCl and NaOH concentrations to achieve pH levels of 3, 5, 7, 9, and 11. Sampling at 5 min intervals up to 120 min was adopted to monitor the tetracycline hydrochloride concentration.

Exploring the Impact of PMS Dosage on TCH Degradation Rate: The optimal catalyst concentration was maintained at 0.5 g/L, with the tetracycline hydrochloride solution concentration set at 20 mg/L. The PMS concentrations were systematically adjusted to 0.05 mM, 0.07 mM, 0.1 mM, 0.4 mM, and 0.7 mM, with the pH set to the optimum level. Sampling occurred at 5 min intervals up to 120 min to assess the tetracycline hydrochloride concentration.

Exploring the Impact of Catalyst Dosage on TCH Degradation Rate: With a PMS concentration of 0.1 mM and tetracycline hydrochloride solution concentration of 20 mg/L, various catalyst concentrations of 0.1 g/L, 0.25 g/L, 0.5 g/L, and 0.75 g/L were tested, adjusting the pH to the optimal level. Sampling occurred at 5 min intervals up to 120 min to monitor the tetracycline hydrochloride concentration.

Exploring the Impact of Tetracycline Hydrochloride Concentration on TCH Degradation Rate: Employing a PMS concentration of 0.1 mM and optimal catalyst concentration of 0.5 g/L, the tetracycline hydrochloride solution concentrations were diversified to 5 mg/L, 10 mg/L, 20 mg/L, and 50 mg/L, with the pH adjusted to the optimal level. Sampling was conducted at 5 min intervals up to 120 min to track the tetracycline hydrochloride concentration.

All degradation experiments were performed in triplicate. Data are presented as mean ± standard deviation. Statistical significance between experimental groups was determined by one-way ANOVA followed by Tukey’s post hoc test (SPSS 26.0, IBM), with p < 0.05 considered significant. Pseudo-first-order rate constants (k_obs) were derived from linear regression of ln(C_t/C₀) versus time.

The calculation methods for the degradation rate and reaction rate constants are provided in the Supplementary Materials (Equations (S2) and (S3)).

2.5. Quenching Experiments

To identify the active radicals responsible for TCH degradation within the reaction system, we conducted electron paramagnetic resonance (EPR) tests using DMPO as a radical capture agent. Furthermore, to elucidate the predominant radicals in the reaction system, we performed radical quenching experiments. Methanol (MeOH) and tert-butanol (TBA) were employed as radical scavengers for SO₄^·- and ·OH, respectively, with TBA effectively quenching ·OH and MeOH effectively quenching both ·OH and SO₄^·-. The PMS concentration was set at 0.1 mM, and the optimal catalyst concentration was determined to be 0.5 g/L. The concentration of the tetracycline hydrochloride solution was maintained at 20 mg/L, with the MeOH and TBA concentrations set at 20 mM. The pH was adjusted to the optimal level, and samples were collected at intervals of 5, 10, 15, 20, 30, 60, 90, and 120 min to monitor the tetracycline hydrochloride concentration.

3. Results and Discussion

3.1. Characterization of CuFe Catalysts

3.1.1. SEM Analysis

Scanning electron microscopy (SEM) was utilized to analyze the surface morphologies of the three catalysts. Figure 1 displays the SEM images and mapping scans of the three catalysts with varying ratios. Figure 1b,f,j correspond to the SEM images of CuFe(2), CuFe(1), and CuFe(3), respectively. From Figure 1b,f,j, it is evident that the surfaces of all three catalysts consist of small particles with irregular chunks attached. A higher iron content results in more flaky structures, whereas increased copper content leads to agglomeration. They exhibit varying degrees of agglomeration, forming irregular block-like agglomerates, which may have resulted from the van der Waals forces, Coulombic forces, and chemical bonding interactions between Cu and Fe. Figure 1c–e,g–i,k–m correspond to the elemental mapping images of O, Cu, and Fe for CuFe(2), CuFe(1), and CuFe(3), respectively, confirming the presence of both Cu and Fe elements on their surfaces and thus verifying the successful preparation of the copper–iron bimetallic catalysts. Specifically, it can be observed from Figure 1b–e that the O, Cu, and Fe elements are uniformly distributed on the catalyst surfaces, which is more conducive to exerting synergistic effects during TCH degradation.

3.1.2. FTIR Analysis

Fourier transform infrared spectroscopy (FTIR) characterized the three prepared catalysts to analyze the functional groups on their surfaces. Figure 2a displays the FTIR spectra of the CuFe(1), CuFe(2), and CuFe(3) catalysts in the range of 400–4000 cm⁻¹. Notably, the absorption peak at 3445.57 cm⁻¹ corresponds to the bending vibration of O-H groups on the catalysts’ surfaces [32], exhibiting a broad spectral feature. Comparative analysis of the three catalysts revealed that CuFe(2) exhibits a relatively larger O-H absorption peak area than the other two catalysts. The post-reaction decrease in the O-H content observed via XPS indicates possible participation of hydroxyl groups in the PMS activation process. The absorption peak at 2351.51 cm⁻¹ is attributed to CO₂, likely introduced from the atmosphere [33]. Furthermore, the peak at 1636.05 cm⁻¹ signifies the stretching vibration of C=O bonds in the hydroxyl groups [34], while the peaks at 637.41 cm⁻¹ [34], 579.59 cm⁻¹, and 521.77 cm⁻¹ correspond to the stretching vibrations of Fe-O bonds [35]. Additionally, the peak at 469.18 cm⁻¹ is associated with the stretching vibration of Cu-O bonds [34]. It is hypothesized that the Fe-O bonds, Cu-O bonds, and their synergistic interaction likely play a primary role in activating PMS. The spectral analysis revealed varying degrees of stretching and shifting in the Fe-O and Cu-O absorption bands across the three catalysts as the Cu/Fe ratio changed. Furthermore, the consistent presence of metal–oxygen bonds in all three catalysts confirmed their successful formation.

3.1.3. XRD Analysis

An X-ray diffraction (XRD) analysis was utilized to characterize the crystal structures of the three prepared catalysts. Figure 2b illustrates the XRD patterns of these catalysts within the 2θ range of 20° to 90°. The CuFe(1) catalyst primarily contained Fe₂O₃ with coexisting Fe₃O₄ and CuO phases. CuFe(2) exhibited an even coexistence of Fe₂O₃, Fe₃O₄, CuO, and Cu₂O. CuFe(3) predominantly contained CuO alongside Fe₂O₃, Fe₃O₄, and Cu₂O. The full peak assignments are provided in Table S1. From the XPS analysis of the best-performing catalyst, CuFe(2), below, it can be seen that the main activators of PMS are Fe²⁺ and Cu²⁺, along with the synergistic effect between them. In contrast, the main component of CuFe(1) is Fe₂O₃, with Fe³⁺ being the primary existing ion. Additionally, the main component of CuFe(3) is Cu²⁺, which can activate PMS but is less effective compared to the synergistic effect of the evenly distributed Cu and Fe bimetals in CuFe(2). This issue similarly exists in CuFe(1).

3.2. Research on Influencing Factors

3.2.1. The Impact of the Reaction System

As depicted in Figure 3a, PMS alone achieved a 34.56% tetracycline hydrochloride degradation rate after 120 min, demonstrating a limited degradation capacity and the challenge of effective removal by PMS alone [36]. In the Cu + PMS system, the degradation of tetracycline hydrochloride by the system reached 65.88% after 120 min of reaction, suggesting that Cu can catalyze the degradation of tetracycline hydrochloride by PMS to some extent.

In contrast, in the CuFe + PMS system, the degradation rates of tetracycline hydrochloride catalyzed by CuFe(1), CuFe(2), and CuFe(3) reached 75.88%, 81.91%, and 76.62%, respectively, after 120 min of reaction, all surpassing the degradation rates of the PMS alone and Cu + PMS systems. This analysis implies a potential synergistic effect between the Fe and Cu on the catalysts’ surfaces, with the bimetallic iron–copper composition providing more catalytic active sites. Both metals can simultaneously activate PMS, leading to the generation of more SO₄^·- and ·OH radicals, thereby accelerating the catalytic reaction [37,38].

Among the three catalysts with varying copper–iron ratios, CuFe(2) demonstrated the highest degradation efficiency. This phenomenon can be explained through SEM and XRD analyses. The SEM images (Figure 1b–e) reveal that the Cu and Fe elements are most uniformly distributed in CuFe(2), whereas CuFe(3) exhibits Cu aggregation, and CuFe(1) contains a higher iron content. The XRD (Figure 2b) further confirms that the primary component of CuFe(1) is Fe₂O₃, where the trivalent iron (Fe³⁺) cannot effectively catalyze SO₄^·- generation [39]. For CuFe(3), the SEM clearly demonstrates a reduced active site density due to the Cu agglomeration caused by the excessive copper content. Furthermore, the reaction rate constant (k_obs) of CuFe(2) was measured at 0.0118 min⁻¹, surpassing that of the other systems (as depicted in Figure 3b). Consequently, the subsequent experiments focused on CuFe(2) for further investigation.

3.2.2. The Impact of pH

Figure 3c illustrates that at pH values of 3, 5, 7, 9, and 11, the degradation rates of tetracycline hydrochloride by the system after 120 min of reaction were 85.29%, 79.12%, 75%, 71.76%, and 56.76%, respectively. Within the pH range of 3 to 11, the degradation rate of tetracycline hydrochloride decreased as the pH value increased, with the highest degradation rate observed at pH = 3. This phenomenon can be attributed to the predominance of SO₄^·- as the main active species within the pH range of 3 to 7. As the pH increases, the concentration of ·OH also increases, particularly when the pH > 10, where the concentration of ·OH significantly rises due to the potential reaction of SO₄^·- with OH^- in the solution, generating ·OH (as described in Equation (1)) [40,41,42]. The redox potential of ·OH is relatively lower than that of SO₄^·-, leading to decreased catalytic effects. Additionally, iron ions can remain stable in acidic solutions; however, at pH > 7, they may form hydroxide precipitates, diminishing the catalyst’s activation performance [43].

$${SO}_4^{·-}+{OH}^{·-}\to {SO}_4^{2-}+·OH$$

(1)

Furthermore, as depicted in Figure 3d, at pH 3, the reaction rate was measured at 0.01151 min⁻¹. Subsequently, as the pH value increased gradually to 11, the reaction rate declined to 0.00429 min⁻¹. It was observed that at pH = 3, the degradation rate of tetracycline hydrochloride by the system peaked, indicating optimal conditions for catalytic activity.

3.2.3. The Impact of PMS Concentration

PMS serves as the reactant that supplies free radicals in the reaction system, and its concentration directly influences the generation of free radicals. As illustrated in Figure 3e, after 120 min of reaction, when the PMS concentrations were 0.05 mM, 0.07 mM, 0.1 mM, and 0.4 mM, the degradation rates of tetracycline hydrochloride were 64.71%, 74.41%, 85.29%, and 94.12%, respectively. The degradation rate of tetracycline hydrochloride increased with a rising PMS concentration, as within a certain range higher PMS concentrations lead to the generation of more free radicals, thereby augmenting the degradation of tetracycline hydrochloride [44]. However, when the PMS concentration reached 0.7 mM, the degradation rate of tetracycline hydrochloride decreased to 90.29% after 120 min of reaction. This decrease in the degradation rate could be attributed to the scavenging effect of HSO₅^-; the excess of HSO₅^- in the system reacted with the existing SO₄^·- to form SO₅^·- (as described in Equation (2)) [45], thereby reducing the degradation rate of tetracycline hydrochloride.

$${HSO}_5^{·-}+{SO}_4^{·-}\to {SO}_4^{2-}+{SO}_5^{·-}+H^{+}$$

(2)

Moreover, as depicted in Figure 3f, as the PMS concentration increased from 0.05 mM to 0.4 mM, the system’s reaction rate rose from 0.00518 min⁻¹ to 0.01748 min⁻¹. However, further increasing the PMS to 0.7 mM decreased the rate to 0.01543 min⁻¹. This trend confirms the PMS concentration’s impact on tetracycline hydrochloride degradation efficiency, consistent with previous discussions.

3.2.4. The Impact of Catalyst Dosage

Figure 4a illustrates the effect of catalyst dosage on tetracycline hydrochloride degradation in the CuFe(2)/PMS system. Increasing the catalyst dosage from 0.1 g/L to 0.5 g/L enhanced the tetracycline hydrochloride removal rate from 77.21% to 85.29%. This improvement was due to the greater availability of surface active sites at higher catalyst loadings, promoting an increased generation of sulfate radicals (SO₄^·-) and hydroxyl radicals (·OH) from the persulfate, thereby enhancing the degradation efficiency [29,46]. However, when the catalyst dosage was further increased to 0.75 g/L, the degradation rate of tetracycline hydrochloride decreased to 70% after 120 min of reaction. This phenomenon could be due to the excess radicals produced by the higher amount of CuFe(2) catalyzing the PMS, leading to self-quenching reactions (as described in Equations (3) and (4)) [29].

$$·OH+{SO}_4^{·-}\to {HSO}_5^{-}$$

(3)

$${SO}_4^{·-}+{SO}_4^{·-}\to S_2{O}_8^{2-}$$

(4)

Moreover, as depicted in Figure 4b, increasing the catalyst dosage from 0.1 g/L to 0.5 g/L raised the system’s reaction rate from 0.01025 min⁻¹ to 0.01151 min⁻¹. However, further increasing the dosage to 0.75 g/L decreased the rate to 0.00586 min⁻¹. This demonstrates an initial increase followed by a decrease in the reaction rate with a varying catalyst dosage, consistent with the earlier experimental findings.

3.2.5. The Impact of TCH Concentration

Observing Figure 4c, it is evident that at concentrations of 5 mg/L, 10 mg/L, 20 mg/L, and 50 mg/L of tetracycline hydrochloride, the degradation rates at 120 min of reaction were 93.53%, 87.94%, 85.29%, and 52.59%, respectively. This trend demonstrates a decrease in the degradation rate with an increasing tetracycline hydrochloride concentration. This phenomenon arises because, with a fixed amount of catalyst and persulfate (PMS), the production of sulfate radicals (SO₄^·-) and hydroxyl radicals (·OH) in the reaction system remains constant. Consequently, when the tetracycline hydrochloride concentration surpasses a certain threshold, it exceeds the degradation capacity of these radicals, leading to a reduction in the degradation efficiency. Furthermore, Figure 4d shows that as the tetracycline hydrochloride concentration increased from 5 mg/L to 50 mg/L, the reaction rate of the system decreased from 0.01527 min⁻¹ to 0.00105 min⁻¹. Hence, a lower concentration of tetracycline hydrochloride is favorable for its degradation.

3.2.6. Reuse of Catalysts

In order to evaluate whether the CuFe(2) catalyst still had good catalytic activity after multiple uses, after the reaction the catalyst was filtered, washed, and dried for the next catalysis. It can be observed from Figure 4e that the removal rate of TCH gradually decreased after four repeated cycle experiments; the removal rate of TCH was 94.12% after the first use and the reaction was 120 min. The removal rate of TCH still reached 81.91% after four cycles, and the decrease in the degradation efficiency may have been due to the slight dissolution of metal ions on the surface of the catalyst [47], which indicates that the CuFe(2) catalyst still had good catalytic activity after repeated use. This lays a good foundation for its application to actual wastewater.

3.3. Identification of Reactive Species

3.3.1. Electron Paramagnetic Resonance (EPR) Analysis

The electron paramagnetic resonance (EPR) test is a valuable method for identifying the types of free radicals involved in catalytic degradation reactions and has seen widespread use. In this experiment, DMPO served as a spin-trapping agent to detect the types of free radicals generated during the catalytic degradation process using EPR technology. Through an analysis of the EPR signals of the spin adducts DMPO-·OH and DMPO-SO₄^·-, formed by their reaction with SO₄^·- and ·OH, the types of free radicals present during the catalytic degradation process were distinctly identified [48].

The experimental spectrum, depicted in Figure 5a, displayed clear detection signals of DMPO-·OH and DMPO-SO₄^·- in the CuFe(2) + PMS system, indicating the presence of SO₄^·- and ·OH as the free radicals in this system. Furthermore, comparison of signal intensities at different time points revealed a stronger detection signal of DMPO-·OH at 10 min compared to 5 min, possibly due to the reaction of SO₄^·- with H₂O to produce ·OH (Equation (5)) [49]. Despite the consumption of SO₄^·- during its conversion to ·OH, the continuous activation of PMS by CuFe(2) helped maintain the concentration of SO₄^·-, resulting in stable signal intensity of DMPO-SO₄^·- in the CuFe(2) + PMS system without significant fluctuations [50].

$${SO}_4^{·-}+H_{2}O\to {HSO}_4^{·-}+·OH+H^{+}$$

(5)

The findings suggest that the radical species within the CuFe(2) + PMS system are SO₄^·- and ·OH.

3.3.2. Free Radical Quenching Experiment

To investigate the copper–iron oxide’s PMS activation mechanism, identifying the dominant radical in the CuFe(2)/PMS system is essential. The electron paramagnetic resonance (EPR) spectroscopy confirmed hydroxyl radicals (·OH) and sulfate radicals (SO₄^·-) as the primary radicals generated. Radical quenching experiments using methanol (MeOH) and tert-butanol (TBA) further identified the key active species, with MeOH scavenging both ·OH and SO₄^·-, while TBA selectively quenched ·OH. As depicted in Figure 5b, the degradation rate of tetracycline hydrochloride without a quencher was 85.29%. Upon adding TBA and MeOH, the degradation rates decreased to 80.88% and 68.68%, respectively, underscoring the primary role of SO₄^·- in tetracycline hydrochloride’s degradation within this reaction system.

3.4. XPS Analysis

X-ray photoelectron spectroscopy (XPS) serves as a critical method for characterizing surface functional groups and valence states in materials. To elucidate the persulfate activation mechanism by the catalyst, XPS analyses were conducted on the CuFe(2) catalyst pre- and post-reaction. Figure 6a presents the overall XPS spectra of CuFe(2) before and after the reaction, indicating consistent detection of Cu and Fe elements on the surface, suggesting no change in the surface elemental composition pre- and post-application.

The O1s peaks of CuFe(2) before and after the reaction were differentiated using XPS peak fitting, as shown in Figure 6b,c. The peak at 529.28 eV corresponds to lattice oxygen (Me=O), 530.7 eV to hydroxyl oxygen (O-H), and 535.00 eV to H₂O peak [51]. Presence of O-H and Me=O peaks indicates oxygen-rich functional groups in CuFe(2). There was a change in the O-H and Me=O peak proportions post-reaction, with the Me=O increasing from 20.09% to 56.12%, while the O-H decreased from 69.74% to 43.88%, suggesting participation of both lattice oxygen (Me=O) and hydroxyl oxygen (O-H) in PMS activation.

Fe2p spectra of CuFe(2) pre- and post-reaction are shown in Figure 6d,e. The peaks near 723 eV and 710 eV can be differentiated into Fe²⁺ and Fe³⁺ characteristic peaks [52,53]. Pre-reaction, Fe²⁺ and Fe³⁺ accounted for 34.08% and 65.92%, respectively. Post-reaction, Fe²⁺ decreased to 28.31%, while Fe³⁺ increased to 71.69%, indicating a valence state change in the iron during the catalytic degradation, suggesting Fe(II)/Fe(III) involvement in PMS activation through redox reactions.

Cu2p spectra of CuFe(2) pre- and post-reaction are depicted in Figure 6f,g. The peaks near 933 eV, 942 eV, and 952 eV can be differentiated into Cu⁺ and Cu²⁺ characteristic peaks [54,55,56]. Pre-reaction, Cu⁺ and Cu²⁺ were 34.08% and 65.92%, respectively. Post-reaction, Cu⁺ increased to 49.82%, while Cu²⁺ decreased to 50.18%, indicating Cu(I)/Cu(II) involvement in PMS activation through a redox cycle.

3.5. Persulfate Activation Mechanism

The research findings suggest that sulfate radicals (SO₄^·-) and hydroxyl radicals (·OH) are the pivotal active species in the CuFe(2) + PMS system for tetracycline hydrochloride degradation. During catalysis, the surface Fe(II) and Cu(II) separately oxidize and reduce to Fe(III) and Cu(I), fostering peroxymonosulfate (PMS) activation and radical generation. This underscores the synergistic Cu-Fe bimetallic effect [46]. Due to Cu and Fe’s high electron donor capacity, the electron transfer between Cu/Fe aids in free radical production [57]. Hence, the persulfate activation mechanism in the CuFe(2) + PMS system can be summarized as follows (Equations (6)–(10)): Initially, HSO₅^- binds with the surface Cu and Fe via hydrogen bonding, leading to redox reactions that yield SO₄^·- and ·OH. Concurrently, Fe(III) oxidizes Cu(I) to Cu(II) until PMS depletion [34]. The activation mechanism is shown in Figure 7.

$$Fe\left(II\right)+{HSO}_5^{-}\to Fe\left(III\right)+{SO}_4^{·-}+{OH}^{-}$$

(6)

$$Cu\left(II\right)+{HSO}_5^{-}\to Cu\left(I\right)+{SO}_5^{·-}+H^{+}$$

(7)

$$Cu\left(I\right)+{HSO}_5^{-}\to Cu\left(II\right)+{SO}_4^{·-}+{OH}^{-}$$

(8)

$${SO}_4^{·-}+H_{2}O\to ·OH+{SO}_4^{2-}+H^{+}$$

(9)

$$Fe\left(III\right)+Cu\left(I\right)\to Fe\left(II\right)+Cu\left(II\right)$$

(10)

3.6. Possible Degradation Pathways of TCH

To delve into the intermediate products and deduce the degradation pathway of tetracycline hydrochloride (TCH), liquid chromatography–mass spectrometry was utilized. The degradation pathway is shown in Figure 8. Initially, TCH decomposes into tetracycline (TC) with an initial m/z value of 445 [35]. TC may undergo degradation via two pathways when subjected to free radicals. Firstly, demethylation converts TC into intermediate P1 (m/z = 427), while dehydroxylation leads to intermediate P6 (m/z = 413) [6]. Subsequently, both P1 and P6 undergo N-demethylation to yield intermediates P2 (m/z = 397) and P7 (m/z = 386), respectively. Under the influence of ·OH and SO₄^·-, compounds P2 and P7 experience oxidation and ring-opening reactions, resulting in the cleavage of C-C bonds and generating intermediates P3 (m/z = 348) and P8 (m/z = 294). As the rings open further, processes like dealkylation, dehydration, and dehydroxylation give rise to intermediates P4 (m/z = 251), P5 (m/z = 212), P9 (m/z = 196), and P10 (m/z = 136) [46]. Finally, intermediate P11 (m/z = 111) is oxidized by ·OH and SO₄^·- to produce H₂O and CO₂.

4. Conclusions

This study empirically demonstrates that the CuFe(2) catalyst (Cu:Fe = 1:2), synthesized through co-precipitation and N₂-atmosphere calcination, achieved 93.53% tetracycline hydrochloride (TCH) degradation within 120 min under the condition of TCH = 5 mg/L. The structural characterization via SEM and XRD confirmed the uniform elemental distribution and coexistence of Cu₂O, CuO, and Fe₂O₃ phases in CuFe(2), while CuFe(1) predominantly formed catalytically inactive Fe₂O₃ and CuFe(3), which suffered from copper agglomeration. The EPR and radical quenching experiments verified SO₄^·- and ·OH as the primary active species. The XPS analysis directly evidenced the redox cycle between Cu⁺/Cu²⁺ (post-reaction Cu⁺ increased to 49.82%) and Fe²⁺/Fe³⁺ (post-reaction Fe²⁺ decreased to 28.31%) during PMS activation. The catalyst retained 81.91% degradation efficiency after four reuse cycles, and the LC-MS analysis identified 11 intermediate products, indicating that the TCH decomposition pathways involved demethylation, dehydroxylation, and ring-opening reactions. These results provide a practical basis for scaling antibiotic wastewater treatment, and the synthesis strategy offers a transferable approach to optimize bimetallic catalysts for persistent pollutant remediation.