Zero-Valent Iron-Supported Magnetic Hydrochar Derived from Kitchen Waste for Efficient Fenton-like Degradation of Tetracycline Hydrochloride

Abstract

In this study, hydrochars loaded with iron species (Fe@HTC and Fe@HTC−T) were prepared by chemical co-precipitation and tubular furnace sintering treatment to develop efficient and sustainable catalysts for antibiotic wastewater treatment, addressing key challenges in sustainable environmental management. The characterization results indicated that iron species loaded on the hydrochars changed from Fe₃O₄ to FeO and then to metallic Fe with the pyrolysis temperature increased from 400 °C to 800 °C. The results of the characterization revealed a phase transition of iron species, confirming the temperature-dependent evolution of catalytic activity. The catalytic performance of the hydrochar composites was evaluated for tetracycline hydrochloride (TC–HCl) degradation via a Fenton-like process. Under optimal conditions (0.2 g/L TC–HCl, 0.1 g/L catalyst, 0.1 mM H₂O₂, pH = 6.86), Fe@HTC−T demonstrated excellent catalytic activity with a removal efficiency of 91.2%. Moreover, Fe@HTC−T exhibited superior stability and low iron leaching rates, attributed to the protective role of the hydrochar matrix. Mechanism research suggested that hydroxyl radicals (•OH) played a dominant role in the degradation process. This study demonstrates the potential of utilizing low-cost and renewable hydrochar materials derived from biomass waste to address industrial challenges in treating high-concentration antibiotic wastewater, offering a sustainable and cost-effective solution with broad applications in environmental remediation.

1. Introduction

In recent years, tetracycline (TC) has found extensive use in human medicine, animal husbandry, and aquaculture due to its effectiveness in treating bacterial infections [1]. However, the abuse of TC has caused many environmental pollution problems. In particular, tetracycline drugs are resistant to degradation, causing bacteria to accumulate over time in the soil and water, which in turn promotes antimicrobial resistance and threatens the ecosystem [2,3]. Therefore, scientists in the fields of chemistry [4,5], environment [6,7], biology [8], and medicine [9] have placed significant emphasis on the removal of TC. Currently, there are several mature methods to remove TC, such as adsorption [10], ion exchange [11], photodegradation [12], advanced oxidative degradation [13], biodegradation methods [14], etc.

Catalytic degradation technology offers distinct advantages over traditional methods. It operates under mild conditions, reducing energy consumption and operational costs. Additionally, the amount of catalyst can be reduced, enhancing economic feasibility and sustainability. With faster reaction rates, catalytic degradation enables the efficient treatment of large volumes of wastewater or gas. These benefits have made it a widely researched solution in environmental remediation.

Recent advancements in catalytic technologies for degrading TC have shown promising results. For instance, Yang et al. [15] explored the use of metal-foam-supported CoMnNi-layered double hydroxide (LDH) catalysts in degrading enrofloxacin in water. Their study revealed that the catalyst, synthesized in situ on nickel foam, effectively activates peroxymonosulfate (PMS) to degrade the antibiotic, achieving a degradation rate of 82.6% under optimal conditions (0.4 g/L catalyst, pH 6.82, 2.0 mM PMS concentration) within 30 min. Similarly, Wang et al. [16] developed a novel Fe-based perovskite catalyst by doping LaFeO₃ with nickel and combining it with metal oxyhydroxide (MeOOH). This combination exhibited an enhanced catalytic performance, leading to a degradation efficiency of 91.84% for ofloxacin within just 15 min. On another note, Aya M. Al-Gariaa and colleagues [17] explored the application of manganese-doped ZnO nanospheres in the photocatalytic degradation of cefotaxime, where the addition of manganese significantly enhanced the material’s photocatalytic efficiency, achieving up to 99% drug degradation within two hours. Correspondingly, Radwan et al. [18] synthesized eco-friendly ZnO/CuO nanocomposites using bee by-products, achieving maximum adsorption of 90.14 mg·g⁻¹ for Congo Red dye, offering a sustainable solution for wastewater treatment. These studies not only demonstrate the potential of catalytic methods in enhancing antibiotic degradation but also emphasize the importance of ongoing research and development in this area.

On the other hand, Fenton- or Fenton-like-based advanced oxidation has been widely used for the degradation of organic contaminants due to its simple operation, high efficiency, and no secondary contamination [19,20]. In the process, hydrogen peroxide (H₂O₂) is decomposed into highly oxidizing active hydroxyl radical (•OH) catalyzed by iron salts or iron (hydr)oxides, and the organic pollutants such as tetracycline can be eliminated efficiently by advanced oxidation [21,22,23,24,25,26]. During the process, it has been reported [27,28,29,30,31] that Fe²⁺ can be oxidized into Fe³⁺, after which Fe³⁺ is further reduced to Fe²⁺, as in Equations (1) and (2). However, the wide application of Fenton oxidation was limited owing to the relatively narrow applicable pH range [32] and/or low Fe(III)/Fe(II) recycling efficiency [33,34,35,36,37]. Zero-valent iron (ZVI) can enhance Fenton-like oxidative systems and is commonly applied for the removal of organic contaminants [38,39,40]. In the process, ZVI can speed up the cycle of Fe³⁺/Fe²⁺ owing to the strong reduction property, thus applying sufficient ferric iron in the system, as shown in Equations (3)–(5). However, there are also some disadvantages in the practical application process of ZVI materials, such as its strong tendency of aggregation and oxidation. In particular, iron particle aggregation can notably reduce the effective surface area, thereby decreasing their catalytic reactivity [41,42,43]. To address this, some studies suggest supporting ZVI on various carriers such as clay ore [44,45], activated carbon [46,47], and biochar material [48,49,50].

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{O}\mathrm{H}}^{-}+•\mathrm{O}\mathrm{H}$$

(1)

$${\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}\mathrm{O}}_2•+{\mathrm{H}}^{+}$$

(2)

$${\mathrm{F}\mathrm{e}}^0+{\mathrm{O}}_2+2{\mathrm{H}}^{+}\to {\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2$$

(3)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{O}\mathrm{H}}^{-}+•\mathrm{O}\mathrm{H}$$

(4)

$${2\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{F}\mathrm{e}}^0\to 3{\mathrm{F}\mathrm{e}}^{2+}$$

(5)

Hydrochars, with biomass as raw materials, have been widely applied in pollutant adsorption, degradation, soil remediation, and other fields [51,52,53]. Studies showed that the dispersivity of ZVI particles can be effectively improved by loading ZVI particles on hydrochars [48,54,55]. However, hydrochars have the disadvantage of underdeveloped pore structures, which can be improved effectively via the calcine process [56,57,58,59]. In recent years, hydrochar materials supported by ZVI have been synthesized in situ by reducing ferric salt during the hydrothermal or calcine process [60]. However, the changes in iron’s valence state and the properties of the composites have not been thoroughly investigated.

The innovation of this study lies in the preparation of zero-valent iron (ZVI)-supported hydrochar composites using hydrothermal, metal salt co-precipitation, and sintering methods. The research investigates the influence of sintering temperature on the oxidation states of iron species and their surface functional groups within the hydrochar composites, aiming to enhance the Fenton-like degradation performance of tetracycline (TC). The surface morphology and physical and chemical properties of carbon materials were characterized by SEM, XRD, FTIR, XPS, and other techniques. The carbon materials were subsequently applied as catalysts in the Fenton-like degradation of tetracycline hydrochloride (TC–HCl). The effects of TC–HCl concentration, catalyst dosage, H₂O₂ concentration, and pH on degradation efficiency were systematically studied. The primary objective is to co-sinter hydrochar with metal oxides, modifying the hydrochar matrix and uniformly loading zero-valent iron into its channels to improve the catalytic activity of the composites in Fenton-like reactions. However, the study found that catalytic efficiency decreased after regeneration, potentially due to the oxidation of zero-valent iron, which alters the oxidation states of the iron species during the catalytic process.

2. Experimental Section

2.1. Materials and Instruments

In this experiment, typical samples were extracted to simulate the composition of Kitchen waste, Shanghai green vegetable leaves rich in cellulose, and steamed buns rich in starch. First, the raw materials were dried under 80 °C for 24 h and smashed by a pulverizer. Then, the powders of vegetable leaves and steamed buns were mixed evenly at a mass ratio of 5:1. All the chemical regents used in this work, methanol (MeOH), tertiary butyl alcohol (TBA), para-benzoquinone (p-BQ), tetracycline hydrochloride (C₂₂H₂₄N₂O₈·HCl, 98% purity), H₂O₂ (30 wt %), thiourea (99% purity), FeSO₄·7H₂O, and FeCl₃·6H₂O, were analytical grade and used without further purification.

2.2. Preparation of HTC

Hydrothermal carbonization experiments were performed in PTEF reactors. Typically, 1.5 g raw powders and 10 mL of deionized water were mixed evenly in a PTEF reactor for 30 min, then sealed in a stainless steel reactor and heated under 200 °C in an oven for 240 min, and naturally cooled down to room temperature. The powders of HTC were filtered and washed thoroughly with deionized water, then dried in a vacuum oven at 60 °C for 24 h.

2.3. Preparation of Fe@HTC

Fe@HTC was prepared by the chemical co-precipitation method. Typically, 1.12 g FeCl₃·6H₂O (4.14 mmol) and 0.58 g FeSO₄·7H₂O (2.09 mmol) were added to a beaker containing 100 mL of distilled water. Then, 0.5 g of HTC powders were added to the beaker, and heated to 80 °C under continuous stirring. Subsequently, the pH of the solution was adjusted to 11 slowly with 1 mol/L NaOH solution under 80 °C and stirred continuously for another 30 min. After cooling, the powders of Fe@HTC were separated by filtration and washed repeatedly to neutral with deionized water, and then dried in an oven under 60 °C for 24 h.

2.4. Preparation of Fe@HTC−T

Briefly, 0.35 g of Fe@HTC was placed into a quartz crucible in a tube furnace, calcined under a continuous N₂ flow at 400, 600, and 800 °C for 2 h with a heating rate of 10 °C/min, and then cooled to room temperature, resulting in samples of Fe@HTC−400, Fe@HTC−600, and Fe@HTC–800, respectively.

2.5. Characterization of Materials

The characteristic peaks of carbon materials and their crystal structures were analyzed by X-ray diffraction (XRD). The X-ray diffraction (XRD) patterns were collected on a Rigaku ULTIMA IV diffractometer equipped with Cu Kα in a 2θ range of 5~90°. Scanning Electron Microscopy (SEM) was used to describe the microstructure and morphology of carbon materials under vacuum. The SEM analysis was conducted using a Hitachi S4800. The functional group species and chemical composition of the carbon material surface were analyzed by Fourier Transform Infrared Spectroscopy (Nicolet NEXUS 670) Madison, WI, USA, using KBr pellets in the range of 500~4000 cm⁻¹. X-ray Photoelectron Spectroscopy (XPS, Thermo Fisher ESCALAB 250Xi) was used to analyze the surface element composition of carbon materials and the semi-quantitative analysis of the element ratio. The Brunauer–Emmett–Teller (BET) method was used to analyze the pore structure of the carbon materials, focusing on their specific surface area and porosity. Pore size distributions were calculated using the Barrett–Joyner–Halenda (BJH) method.

2.6. Catalytic Degradation Experiments

The catalytic performance of the obtained samples was analyzed through TC–HCl degradation in a Fenton-like process. Batch experiments were conducted in a shaker in the dark with a fixed speed of 150 r/min. Firstly, the catalytic properties of Fe@HTC and Fe@HTC−T were compared. TC–HCl degradation in different systems was examined under conditions of 100 mL of 0.2 g/L TC–HCl, 0.1 g/L Fe@HTC–800, 0.1 mM H₂O₂, 30 °C. To obtain the optimal experiment parameters for TC–HCl removal, the influences of pH (2~10) and dosages of Fe@HTC–800 (0~0.2 g/L) and H₂O₂ (0~1 mM) were investigated separately.

In the process of each catalytic experiment, TC–HCl aqueous solution containing synthesized catalysts was sonicated for 3 min and transferred to the constant temperature oscillator. After 30 min of adsorption, a specific amount of H₂O₂ was added. After that, 2 mL of the sample was collected at different intervals, filtered through 0.22 μm Millipore membranes, and then mixed with 2 mL of methanol to terminate the reaction immediately. The residual TC–HCl concentration in the reaction solution was then measured using a UV–vis spectrophotometer. The concentration of leached iron in solution was measured by ICP-MS. The solid phase was also filtered and characterized after treatment.

The removal efficiency (R) of TC–HCl was calculated using the following formula:

$$\mathrm{R}(\mathrm{\%})={\displaystyle \frac{{\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_0}}\times 100\mathrm{\%}$$

(6)

where C₀ (g/L) represents the initial concentration of TC–HCl and C_t (g/L) denotes the concentration at reaction time t (min).

Based on the catalyst preparation steps and degradation process outlined above, we have developed a corresponding flowchart, as shown in Figure 1.

3. Results and Discussion

3.1. Characterization of Synthesized Catalyst

Figure 2 shows the XRD patterns of the obtained samples. As shown in Figure 2a, the wide peak located between (2θ) 15~30° for HTC could be attributed to cellulose and amorphous carbon. As shown in Figure 2b, the diffraction peaks of Fe@HTC and Fe@HTC−400 at (2θ) 35.42°, 43.05°, and 62.5° might be ascribed to Fe₃O₄ (JCPDS No. 19-0629). The diffraction peaks of Fe@HTC−600 at (2θ) 36.04°and 41.92° indicated the formation of the FeO (JCPDS No.06-0615) structure. As for Fe@HTC–800, the new peaks at (2θ) 44.8° and 65.2° could be ascribed to the diffraction peak of α-Fe⁰ (JCPDS No.06-0696). The XRD patterns showed that the reduction of iron precursor occurred during the calcination treatment. With the pyrolysis temperature improved from 400 °C to 800 °C, the iron species changed from Fe₃O₄ to FeO and then to metallic Fe. Moreover, the diffraction peak at 21.89° basically disappeared, indicating that the carbon material was further carbonized.

The SEM images showed the surface morphologies of HTC, Fe@HTC, and Fe@HTC–800 (Figure 3 and Figure S1). Fe@HTC was composed of uniform microspheres with sizes in the range of 4~5 μm. After the co-precipitation process, numerous irregular nano-particles were evenly distributed across the sphere surfaces (Figure 3a,b and Figure S1). After calcination at 800 °C for 2 h, the surface of Fe@HTC–800 became obviously rough, and most of the spheres were completely destroyed (Figure 3c,d).

The N₂ adsorption–desorption isotherms and the pore size distribution of Fe@HTC and Fe@HTC–800 were represented in Figure S2. According to the classification of IUPAC, the isotherm of Fe@HTC showed the shape of type IV with distinct hysteresis loops at high pressure, indicating the presence of a mesoporous structure. As for Fe@HTC–800, the isotherm presented a typical type II curve with an H4-shaped hysteresis loop, which was characteristic of the physical adsorption process. Further analyses for the specific surface area and total volume of pores for the synthesized materials are listed in

Table 1. Structural analyses of the synthesized materials.

Sample	SBET (m2/g)	V (cm3/g)	DBJH (nm)
Fe@HTC	60.55	0.170	9.086
Fe@HTC–800	91.78	0.065	3.786

. The surface area of the composite increased from 60.55 to 91.78 m²/g, while its total pore volume reduced from 0.17 to 0.065 cm³/g and its pore size decreased from 9.9 to 3.7 nm. This might be associated with the impregnation of Fe nanoparticles into the framework of HTC as well as the improved ratio of the Fe phase with the decomposition of HTC.

The FT-IR spectra of HTC, Fe@HTC, and Fe@HTC–800 were recorded to analyze the surface functional groups, and the results are shown in Figure 4. The high absorption peaks of HTC and Fe@HTC at 3400 cm⁻¹ and 2900 cm⁻¹ indicated that there are abundant -OH functional groups and saturated C-H bonds on the hydrothermal carbon, while the peak at 1561 cm⁻¹ corresponds to the C=O stretching vibration of the carbonyl or carboxyl groups. However, all of these peaks disappeared or decreased on a large scale for Fe@HTC–800, which means that the functional groups were broken under calcination at 800 °C.

The XPS was performed to analyze the composition and chemical states of Fe@HTC and Fe@HTC–800 (Figure 5). The XPS characterization of Fe@HTC–800−spent is shown in Figure S3. From the full scan spectra (Figure 5a), Fe, C, and O elements could be detected on the surface of Fe@HTC and Fe@HTC–800. Figure 5b shows the C 1s spectrum with the binding energies of 284.73 eV, 286.52 eV, and 288.34 eV, contributing to C=C/C-C, C-O, and C=O, respectively. Figure 5c shows the O 1s spectrum; the peak at 531.81 eV is attributed to the lattice oxygen of the metal oxide (O²⁻). The spectra of C 1s and O 1s indicated the presence of biomolecules, primarily amorphous carbon in hydrothermal carbon covering the surface of the ZVI particles. As shown in Figure 5d, the Fe 2p_3/2 peaks of Fe@HTC at 710.8 eV and 712.0 eV were attributed to Fe²⁺ and Fe³⁺, respectively. After sintering at 800 °C, a new peak at around 706.8 eV can be assigned to Fe⁰. Simultaneously, the peak of Fe³⁺ at 712.0 ev basically disappeared, which confirmed that the Fe species were reduced under calcination. Due to the strong reactivity of the ZVI particles, Fe⁰ is readily oxidized during preparation and storage, forming an iron oxide layer on its surface [57]. The presence of surface materials (amorphous carbon and iron oxides) hinders the detection of Fe⁰ in ZVI particles.

3.2. TC–HCl Removal Using the Synthesized Catalyst

Figure 6a shows the catalytic degradation effects of different systems. TC–HCl was stable over the experimental time and did not degrade autonomously. In the presence of H₂O₂, 34.8% of the TC–HCl was removed within 120 min, indicating that H₂O₂ is not easy to oxidize TC–HCl. In addition, Fe@HTC–800 showed a certain adsorption property, and 58.6% of the TC–HCl was removed within 120 min. When Fe@HTC–800 and H₂O₂ were introduced into the reaction system simultaneously, TC–HCl removal efficiency was significantly improved and it reached up to 91.2%. Moreover, the effects of calcination temperature on the removal capacities of Fe@HTC−T have been studied (Figure 6b). As can be seen, Fe@HTC−T showed a different effect on TC–HCl degradation in the presence of H₂O₂. The removal efficiency of Fe@HTC was only 51.4% after 120 min. With the calcination temperature increase from 400 °C to 800 °C, the removal efficiency was improved obviously from 62.1% to 91.2%. As expected, HTC supported with zero valence iron (Fe@HTC–800) showed the best TC–HCl removal efficiency. This above research suggests that Fe@HTC–800 is a promising effective catalyst for the degradation of TC–HCl in the Fenton-like process.

To further explore the optimal conditions for the catalyst to remove TC–HCl, the effects of several key parameters including the catalyst dosage, H₂O₂ concentration, initial pH, and inorganic anions on TC–HCl removal were systematically evaluated.

The influence of catalyst dosage on TC–HCl removal efficiency is shown in Figure 7a. As observed, the removal efficiency of TC–HCl increased from 78.2% to 93.6% as the catalyst dosage was raised from 0.05 g/L to 0.2 g/L. This enhancement can be attributed to the fact that catalytic activity is directly linked to the number of active sites; a higher catalyst dosage increases the production of Fe²⁺ from zero-valent iron. Fe²⁺ then reacts with H₂O₂ to generate additional hydroxyl radicals (•OH), which have a strong oxidizing power for TC–HCl. However, as the Fe@HTC–800 dosage increased from 0.1 g/L to 0.2 g/L, the TC–HCl removal rates at 120 min became nearly identical. It was reported [58] that excessive catalysts may cause particle agglomeration and lower radical generation. Considering both catalytic efficiency and cost, 0.1 g/L Fe@HTC–800 was regarded as the appropriate dosage for the subsequent experiments.

Previous studies indicated that hydroxyl radicals are mainly produced by H₂O₂ in the Fenton-like system, which will directly affect the degradation of TC–HCl [25,26]. Here, dosages of H₂O₂ on the effect of removal efficiency of TC–HCl was explored. Figure 7b shows that the removal efficiency of TC–HCl increased from 88.7% to 91.2% as the H₂O₂ concentration increased from 0.05 mM to 1 mM. An insufficient amount of H₂O₂ caused the limited generation of hydroxyl radicals. As the concentration of H₂O₂ increased, more hydroxyl radicals were generated and the removal efficiency of TC–HCl was enhanced. However, the removal efficiency of TC–HCl was only slightly improved when the concentration of H₂O₂ increased from 0.1 mM to 1 mM. This is probably because excessive H₂O₂ might compete with TC–HCl for •OH and induce the undesirable decomposition of H₂O₂ [61]. Therefore, the optimum dosage of H₂O₂ for the subsequent experiments was determined as 0.1 mM.

It is well established that the initial pH significantly influences the Fenton-like reaction. As shown in Figure 7c, the removal efficiency of TC–HCl in this study varied evidently over a wide pH range from 2 to 10. When the solution pH increased from 2.0 to 6.86, the TC–HCl removal efficiency at 120 min increased from 69.4% to 91.2%. Then, TC–HCl removal efficiency decreased from 91.2% to 64.9% with pH changed from 6.86 to 10.7. This is probably due to the following three reasons: Firstly, the surface of the ZVI particles is more susceptible to corrosion under acidic conditions, resulting in the release of a significant amount of Fe²⁺ [2,3]. Fe²⁺ reacts with hydrogen peroxide to generate additional hydroxyl radicals, facilitating the oxidation of TC–HCl. With increasing pH, ZVI tends to form a passive oxide layer, blocking the active site of the reaction and thus reducing the degradation efficiency [62]. Secondly, hydrogen peroxide readily decomposes into oxygen and water under alkaline conditions. However, it facilitates the release of hydrogen when the pH is too low. Thirdly, pH is related to the surface charge of the catalyst and TC–HCl. The analysis and comparison of the Zeta potential test between Fe@HTC–800 and TC at pH 2.00–10.00 are presented in Table S1. As shown in Figure S4, Fe@HTC–800 is positively charged on the surface between pH = 2~3.4, while TC–HCl is positively charged between pH = 2~8.9. The catalyst and the substrate are both positively charged at pH = 2, which is against the pre-concentration process. Between pH = 4~8.3, Fe@HTC–800 had negative zeta potential, it adsorbed the positively charged TC–HCl better and provided higher efficiency in the Fenton-like reaction. When pH > 8.28, the pre-enrichment effect is not conducive owing to the negative charge on the catalyst and the substrate. Therefore, pH = 4~8 is more suitable for the catalytic reaction system.

Figure S5 shows the repeatability of the final degradation rate and the corresponding error bar analysis under different initial pH value conditions. According to the bar chart, it can be found that the repeatability of each group of experiments of carbon material can be consistent within the error range, indicating that the degradation effect of the catalyst is relatively reproducible.

In natural water bodies and typical industrial wastewater, certain concentrations of inorganic anions are often present. Previous studies have shown that the presence of anions can influence both the decomposition of H₂O₂ and the degradation of organic compounds during the Fenton-like reaction [63,64]. Based on this, several typical inorganic anions and water from Cherry River at East China Normal University were selected to explore the impact of anionic interference on the Fenton-like reaction. As shown in Figure 7d, the catalyst maintained excellent TC–HCl removal efficiency in the presence of different anions, indicating that the material possessed strong resistance to most of the anionic interference.

The iron leaching rates at different pHs were analyzed using ICP. The leaching rate formula is provided as described in Equation (S1). As shown in Table S2, the leaching rates decreased as the pH increased, which is consistent with the mechanism of the Fenton-like reaction. The leaching of iron was higher under acidic conditions, likely because the acidic environment promoted the oxidation and dissolution of zero-valent iron, thus producing more Fe²⁺ and accelerating the Fenton reaction. However, the leaching rates of iron were low under alkaline conditions, probably due to the involvement of other reaction pathways such as the hydroxide precipitation of iron, which might affect the activity and reaction mechanism of iron. According to the removal efficiency and loss of the catalyst in the reaction system, the range of pH = 4~8 is more appropriate.

3.3. Kinetic Analysis of TC–HCl Degradation

To further understand the degradation process of tetracycline hydrochloride (TC–HCl), kinetic studies were conducted under various conditions. The degradation data were analyzed using the first-order kinetic model, as represented by Equation (7):

$$\mathrm{l}\mathrm{n}\left({\displaystyle \frac{{\mathrm{C}}_0}{{\mathrm{C}}_{\mathrm{t}}}}\right)=-\mathrm{k}\mathrm{t}$$

(7)

where C_t (mg/L) and C₀ (mg/L) represent the concentration of TC–HCl at reaction time t (min) and the initial concentration, respectively, and k is the rate constant.

Figure S6 shows the kinetic analysis of TC–HCl degradation under different conditions. As shown in Figure S6a, the data demonstrate a linear relationship between ln(C_t/C₀) and time, with varying slopes corresponding to different rate constants (k) for each catalyst concentration. As the catalyst concentration increased, the degradation rate accelerated, reflected by the higher values of k. As shown in Figure S6b, an increase in H₂O₂ concentration resulted in a higher rate constant, indicating that higher H₂O₂ concentrations accelerate the oxidation process and enhance the degradation of TC–HCl. As shown in Figure S6c, the rate constants varied with pH, and the highest degradation rate was observed at pH 6.86, suggesting that pH plays a significant role in the degradation process, likely by influencing the formation and stability of reactive species.

The calculated rate constants (k) and correlation coefficients (R²) from all the experiments indicate that the degradation of TC–HCl follows a first-order kinetic model, and the reaction rate is influenced by both the catalyst concentration, H₂O₂ concentration, and pH conditions. These findings provide a deeper understanding of the reaction mechanisms and highlight the factors that control the degradation efficiency of TC–HCl.

3.4. Comparison of Tetracycline Removal Efficiency Using Different Catalysts

This study compared the TC–HCl removal efficiency of the synthesized composites with other relevant catalysts reported in the literature, and the results are presented in

Table 2. Comparison of tetracycline removal efficiency using various catalysts.

Catalyst	Catalyst Dosage (g/L)	H2O2 (mM)	TC (mg/L)	pH	Removal Efficiency (%)	Reference
Fe@HTC–800	0.1	1	200	6.86	91.2%	This work
BS−ZVI@MB	1.2	50	3	7.2	100%	[65]
nZVI/SEP	1.0	1	20	7	92.6%	[66]
PVP−K30	0.1	-	100	6.5	98.4%	[67]
Fe3O4 nanospheres	0.5	50	25	7	82%	[68]
Fe−GAC	0.4	-	10	2–5	89.4%	[26]
Fe/SCN	0.5	80	10	7	90.3%	[69]

. It can be observed that the synthesized catalyst showed good catalytic performance for high concentrations of TC–HCl, even with low dosages of catalyst and H₂O₂. Moreover, the raw materials were inexpensive, readily available, and non-toxic. At the same time, the composite catalyst preparation process was simple and easy to replicate. Thus, the synthesized composite is anticipated to be an effective catalyst for TC–HCl degradation in aqueous solutions.

4. Proposed Degradation Mechanism of TC–HCl by Fe@HTC–800 Composite

The contribution of reactive oxygen species (ROS) was explored by a quenching experiment. Here, MeOH and TBA were employed as the scavengers to •OH, and p-BQ was used to quench •${\mathrm{O}}_2^{-}$. As shown in Figure 8, the removal efficiency of TC–HCl just decreased slightly to 80.58% after the addition of p-BQ, suggesting that •${\mathrm{O}}_2^{-}$ had little effect on TC–HCl degradation. However, when MeOH and TBA were added, the degradation efficiency decreased sharply from 91.23% to 46.93% and 54.15%, indicating that •OH radicals played a key role in the degradation process.

In the Fenton-like reaction, Fe⁰ exhibits strong reductive properties, allowing it to react with water or H⁺ in the reaction system to generate Fe²⁺. The addition of H₂O₂ promotes the corrosion of Fe⁰, leading to the continuous production of a large amount of Fe²⁺. Subsequently, Fe²⁺ reacts with H₂O₂ to produce •OH radicals, which are highly oxidative and serve as the main reactive species for the degradation of TC–HCl. Additionally, the presence of Fe⁰ accelerates the reduction of Fe³⁺ back to Fe²⁺, further sustaining the reaction. The Fe²⁺/Fe³⁺ redox cycle also promotes the generation of •HO₂ and •O₂⁻ radicals, which contribute to the degradation process. The reaction mechanisms are shown in Equations (8)–(12). Throughout this process, the hydrochar plays a crucial role due to its porous structure. It effectively adsorbs TC–HCl molecules and pre-concentrates them near the catalyst surface. This enhances the direct interaction between TC–HCl and the catalyst, which enhances the subsequent oxidative degradation reactions.

$${\mathrm{F}\mathrm{e}}^0+{\mathrm{H}}_2{\mathrm{O}}_2+2{\mathrm{H}}^{+}\to {\mathrm{F}\mathrm{e}}^{2+}+2{\mathrm{H}}_2\mathrm{O}$$

(8)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{O}\mathrm{H}}^{-}+•\mathrm{O}\mathrm{H}$$

(9)

$${2\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{F}\mathrm{e}}^0\to 3{\mathrm{F}\mathrm{e}}^{2+}$$

(10)

$${\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{2+}+•\mathrm{H}{\mathrm{O}}_2+{\mathrm{H}}^{+}$$

(11)

$${\mathrm{F}\mathrm{e}}^{3+}+•\mathrm{H}{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{O}}_2+{\mathrm{H}}^{+}$$

(12)

To further investigate the degradation mechanism of tetracycline, the intermediate products in the reaction solution were analyzed. Figure 9 shows several possible pathways for TC–HCl degradation in the Fenton-like system, inferred from the literature [70,71,72]. Pathway I involves the cleavage of functional groups and ring structures, producing intermediates such as T1 (m/z = 389) and T2 (m/z = 346), which undergo further degradation into smaller molecules like T4 (m/z = 258). Pathway II predominantly involves hydroxylation and gradual fragmentation of the tetracycline backbone, resulting in intermediates such as T5 (m/z = 446) and T8 (m/z = 114), which are more susceptible to biodegradation. Pathways III and IV involve deamination, hydroxylation, and ring-opening reactions, forming intermediates like T10 (m/z = 463) and T15 (m/z = 149), indicating the effective mineralization of TC–HCl. In Pathway IV, TC–HCl undergoes further decomposition, including deamination, ring opening, and hydroxylation, leading to the formation of smaller molecular species. These smaller intermediates are generally more biodegradable and are progressively degraded into low-molecular-weight products. A significant portion of these intermediates undergo mineralization, ultimately converting into carbon dioxide and water.

5. Conclusions

In this study, hydrochars loaded with iron species (Fe@HTC and Fe@HTC−T) were successfully prepared through chemical co-precipitation and tubular furnace sintering treatment. The characterization results indicated that iron species loaded on the hydrochars changed from Fe₃O₄ to FeO and then to metallic Fe, with an increase in pyrolysis temperature from 400 °C to 800 °C. The Fe@HTC–800 composites exhibited excellent degradation efficiency for TC–HCl in an aqueous system, with low dosages of catalyst and H₂O₂. The mechanism studies revealed that •OH radicals played a key role in the degradation process. The findings of this study demonstrate that Fe@HTC–800 composites hold significant potential for the remediation of TC–HCl wastewater. However, the study found that the catalytic efficiency decreased after the material regeneration, which could be attributed to the oxidation of zero-valent iron during the catalytic process, leading to an increase in its oxidation state. Overall, this research emphasizes that zero-valent iron-based hydrochar composites provide a cost-effective and sustainable solution for antibiotic wastewater treatment, offering valuable insights into the application of hydrochar materials in environmental remediation.