A Novel Double-Coated Persulfate Slow-Release Material: Preparation and Application for the Removal of Antibiotics from Groundwater

Abstract

Single-layer slow-release materials have short lifespans due to their rapid initial release behavior. To address this problem, a double-coated persulfate slow-release material was developed in this study. The outer coating layer consists of polycaprolactone–silica sand, which is used to encapsulate an inner layer of polycaprolactone–silica sand and sodium persulfate. Static and dynamic release experiments were conducted to analyze the behavior and degradation capabilities of this material when activated by iron–nitrogen co-doped biochar (Fe@N-BC) for the removal of sulfamethoxazole (SMZ) and ciprofloxacin (CIP) in groundwater. The double-coated material maintains a stable release rate, achieving optimal performance with an outer layer thickness of 0.25 cm and a silica sand to polycaprolactone (PCL) mass ratio between 2 and 5. Optimal degradation rates for SMZ and CIP were observed at a pH of 3. Specifically, 1 mg/L of SMZ was fully degraded within 12 h, while the complete removal of 1 mg/L of CIP occurred within just 2 h. The presence of humic acid and higher initial pollutant concentrations reduced the degradation rates. Among the tested anions, HCO₃⁻ had the most significant inhibitory impact, while Cl⁻ had the least significant impact on degradation performance. Column experiments demonstrated a consistent release of persulfate over a period of 60 days at a flow rate of 0.5 mL/min. Increased flow rates resulted in a shorter lifespan for this slow-release material. The minimum outflows of SMZ and CIP were obtained with a quartz sand mesh size of 40–60 and a flow rate of 0.5 mL/min. These results offer a theoretical basis for the prolonged and stable release of persulfate, as well as the efficient removal of SMZ and CIP from groundwater.

1. Introduction

Groundwater is a crucial freshwater resource that is essential for ecological balance, food safety, and socio-economic development [1]. However, the widespread use of antibiotics in medicine and livestock farming has led to significant antibiotic contamination in groundwater through various pathways, such as agricultural activities and wastewater discharge [2]. Eighty-four groundwater samples were collected in the North China Plain of China, and 12 antibiotics were detected, with total concentrations ranging from 5.33 ng/L to 64.73 ng/L. CIP and SMZ are the two main antibiotics [3]. This contamination poses a global environmental crisis that threatens aquatic ecosystems and human health by promoting antibiotic resistance [4]. Therefore, strengthening the monitoring and management of antibiotic pollution in groundwater and finding effective, lasting purification technologies are crucial for protecting groundwater resources and ensuring the sustainable development of human society.

In situ chemical oxidation (ISCO) technology serves as a crucial approach for addressing organic contamination in groundwater. ISCO techniques employ oxidants, including Fenton’s reagent, potassium permanganate, and persulfate, to effectively eliminate pollutants [5,6]. Recent years have seen extensive research and application of persulfate in the in situ chemical oxidation remediation of groundwater contaminated with organic pollutants. The standard reduction potential of PDS stands at 2.01 V, surpassing that of hydrogen peroxide (1.77 V) and potassium permanganate (1.67 V), yet it is slightly lower than that of ·OH (2.07 V) and ozone (2.07 V). While certain organic compounds can directly react with persulfate, these reactions are relatively slow. Thus, activation is usually required to efficiently remove pollutants. The activation of persulfate causes the formation of sulfate radicals (SO₄⁻), which have an oxidation–reduction potential of 2.6 V. Thus, a greater potential to degrade a wider variety of pollutants is provided compared to persulfate before activation (2.01 V) [7]. Typical methods for persulfate activation include thermal activation, transition metal activation, alkaline activation, carbon material activation, UV light, cavitation, zero valent iron (ZVI), photocatalysis, etc.

Traditional ISCO technology involves the injection of oxidants into aquifers via gravity pumps, either vertically or horizontally, or through pressurized direct push methods [8,9,10]. However, this approach faces challenges such as oxidant backflow, poor long-term effectiveness, and pollutant rebound [11]. Therefore, a major challenge in successfully implementing ISCO for groundwater remediation is ensuring the continuous and slow delivery of oxidants to the contamination zone. This will enable the long-term presence of oxidants in the groundwater, mitigating pollutant rebound [12]. The emergence of slow-release technology offers a new direction for solving these challenges, and the use of slow-release oxidant materials has emerged as a novel method for the ISCO of groundwater. Controlling the release of oxidants from slow-release materials ensures the sustained delivery of oxidants and that sufficient quantities of oxidants are provided to participate in the reaction [13].

The use of persulfate slow-release oxidizing materials for groundwater organic pollution remediation has been thoroughly investigated (summarized in Table S1) [14,15,16,17,18,19,20,21]. Xu et al. employed sol-gel and gelatin as structural materials to develop persulfate gel slow-release and persulfate gelatin slow-release materials, which, respectively, achieved 78.6% and 66.9% removal efficiencies within 3 h for the removal of 2,4-dinitrotoluene. In this reaction, the persulfate oxidatively degrades the 2,4-dinitrotoluene through a non-radical pathway without the addition of catalysts [22]. Evans et al. used paraffin-encapsulated sodium persulfate to degrade 1,4-dioxane and chlorinated solvents, achieving a removal efficiency of up to 99% over 208 days. This persulfate slow-release agent is superior to permanganate slow-release agents for the degradation of dioxane [23]. Pham et al. used zeolites, diatomaceous earth, and silica to prepare granulated persulfate slow-release materials. These materials steadily released persulfate, achieving a 99% degradation efficiency for a 15 mg/L concentration of trichloroethylene [24]. Tang et al. introduced the innovative use of chitosan and urea as encapsulating agents to create a novel persulfate slow-release material aimed at degrading methyl orange and pyrene. Their findings revealed that urea was essential in the persulfate release mechanism and could additionally catalyze persulfate activation via Fe, leading to the generation of free radicals and an increased degradation rate of methyl orange [25]. T. Wang et al. used O-MnOx and PDS to prepare slow-release materials, which were used to degrade tetracycline in a simulated groundwater environment. This study also explored the optimal ratio of O-MnOx and PDS used to prepare slow-release materials [26].

The scaffold material utilized in the preparation of a slow-release material is pivotal in regulating the controlled release of the oxidant. Paraffin is one of the most commonly used scaffold materials [27]. Chokejaroenrat et al. created a persulfate slow-release material by combining paraffin, persulfate, and zero-valent iron. The utilization of this slow-release material does not require the addition of oxidants or activators, making it suitable for the long-term treatment of organic pollutants in water [28]. J. Ma et al. fabricated slow-release bead materials using persulfate and paraffin, and batch experiments revealed that the bead size and temperature significantly affected their persulfate release kinetics [29]. Zhu et al. conducted column experiments on slow-release materials prepared using paraffin and sodium persulfate. This work enhances the understanding of the release characteristics and dynamics of persulfate slow-release systems [30]. However, further research has gradually revealed the shortcomings of paraffin as a scaffold material, such as its solubility in organic pollutants [31]. Additionally, paraffin may react with sulfate radical anions, which leads to the undesirable consumption of persulfate and affects its effective release and oxidative performance [23]. These challenges point to several directions for future research: (1) New slow-release material scaffolds should be developed by searching for environmentally friendly and effective new slow-release materials that can control the rate of persulfate release to enhance remediation efficiency and environmental adaptability. (2) Long-term effect studies should be performed by extending the experimental degradation period and conducting long-term monitoring studies to more accurately assess the durability and stability of slow-release materials. (3) Field application research should be performed by testing the effectiveness of persulfate slow-release materials under simulated actual groundwater remediation conditions, and various groundwater environments should be evaluated. (4) The synchronous release of persulfate activators should be evaluated by studying synchronous release strategies and their impact on pollutant removal efficiency. An in-depth exploration of these research directions is expected to promote the development of persulfate slow-release technology, providing more effective and environmentally friendly solutions for the treatment of persistent organic pollutants in groundwater.

Most existing slow-release materials also suffer from additional issues that limit their performance. For instance, slow-release materials typically exhibit excessively high oxidant release rates in the initial stage of application, followed by insufficient release later in the degradation process. This leads to a shortened lifespan. Moreover, material deformation during use causes oxidant leaks. Therefore, the scaffold materials and fabrication processes of slow-release materials must be further optimized [32,33]. Studies have shown that wrapping persulfate slow-release materials with an additional layer that exhibits different permeability can address deformation issues while simultaneously achieving the stable release of persulfate. This extends the lifespan of these slow-release materials and improves their utilization [34,35]. In this paper, a double-coated persulfate slow-release material was prepared using melt injection molding. Polycaprolactone (PCL) was employed as the scaffold material, and silica sand was utilized as the porogen. Static release tests were performed to examine how the composition ratio of the inner and outer layers, as well as the thickness of the outer layer, influence the release behavior of persulfate. Using sulfamethoxazole (SMZ) and ciprofloxacin (CIP) as target pollutants, batch degradation experiments were performed to evaluate the static degradation effects of this double-coated persulfate slow-release material. Then, a one-dimensional sand column simulation experiment was conducted to analyze the release characteristics of the double-coated persulfate slow-release material under dynamic conditions and study its feasibility for groundwater pollution remediation.

2. Materials and Methods

2.1. Materials

The materials used in this study include: PDS (Na₂S₂O₈, Tianjin Damao Chemical Reagent Factory, Tianjin, China), potassium iodide (KI, Tianjin Fuyu Fine Chemical Co., Ltd., Tianjin, China), sodium bicarbonate (NaHCO₃, Tianjin Damao Chemical Reagent Factory, China), trisodium phosphate (Na₃PO₄·7H₂O, Tianjin Damao Chemical Reagent Factory), silicon dioxide (SiO₂, Tianjin Damao Chemical Reagent Factory), polycaprolactone (PCL, Tianjin Fuyu Fine Chemical Co., Ltd.), methanol (MeOH, Tianjin Fuyu Fine Chemical Co., Ltd.), ethanol (EtOH, Tianjin Fuyu Fine Chemical Co., Ltd.), furfuryl alcohol (C₅H₆O₂, Tianjin Fuyu Fine Chemical Co., Ltd.), SMZ (C₁₀H₁₁N₃O₃S,98%, McLean Reagent Company, Shanghai, China), CIP (C₁₇H₁₈FN₃₀N,98%, McLean Reagent Company), acetic acid (CH₃COOH), acetonitrile (CH₃CN, McLean Reagent Company), ferrous sulfate heptahydrate (FeSO₄·7H₂O, Tianjin Damao Chemical Reagent Factory), urea (CH₄N₂O, Tianjin Damao Chemical Reagent Factory), ascorbic acid (C₆H₈O₆, Tianjin Damao Chemical Reagent Factory), sodium chloride (NaCl, Tianjin Damao Chemical Reagent Factory), sodium hydroxide (NaOH, Xi’an Chemical Reagent Factory, Xi’an, China), phosphoric acid (H₃PO₄,85%, Xi’an Chemical Reagent Factory), hydrochloric acid (HCl, Tianjin Fuyu Fine Chemical Co., Ltd. China), humic acid (98%, Tianjin Fuyu Fine Chemical Co., Ltd.), sodium nitrate (NaNO₃, Xi’an Chemical Reagent Factory). All the water used in this experiment was ultrapure water, with a resistivity of 18 M Ω∗cm.

2.2. Preparation of Slow-Release Materials

First, a certain amount of PCL was weighed in a beaker and heated to 70 °C under constant -temperature magnetic stirring. Upon the complete melting of PCL, pre-weighed silica sand was incorporated, and the mixture was thoroughly stirred using a glass rod. Subsequently, a specified quantity of Na₂S₂O₈ powder was introduced, and the resulting blend was thoroughly mixed. While still hot, the mixture was poured into a mold. Following the cooling process to room temperature and subsequent solidification.

The outer layer consisted of a coating prepared using PCL and silica sand in various ratios. The thickness of the outer material was controlled using cube molds with different side lengths. Firstly, mix silica sand and PCL in the desired ratio, maintain the mixture at 70 °C, and stir magnetically. The obtained mixture was used to encase the previously obtained inner layer in a mold, and the resulting double-coated persulfate slow-release material was allowed to solidify, as shown in Figure S1a. The ratio of the double-coated slow-release material and the thickness of the outer layer are depicted in Figure 2a.

2.3. Preparation of Iron–Nitrogen Co-Doped Biochar (Fe@N-BC)

Ferrous sulfate heptahydrate, urea, and ascorbic acid were dissolved in deionized water. Subsequently, lotus leaf powder was introduced into the solution and allowed to react fully. Post-reaction, the resulting material underwent centrifugation and vacuum freeze-drying and was then subjected to firing in a tube furnace at 900 °C for 2 h, yielding iron–nitrogen co-doped biochar (Fe@N-BC). The mass-to-volume ratio of the lotus leaf powder, ferrous sulfate heptahydrate, urea, ascorbic acid, and deionized water was 3 g:5.56 g:3 g:2.5 g:100 mL [36]. (The effect of biochar modification on antibiotic degradation is shown in Figure S2).

2.4. Slow-Release and Degradation Performance Testing

2.4.1. Static Release Experiment

The prepared persulfate slow-release materials were added to conical flasks, each containing 200 mL of ultrapure water. The flasks were then sealed, and the persulfate was gradually released at room temperature (25 ± 1 °C) for 21 days. At various time intervals, 0.1 mL of leachate was collected from each flask, and the concentration of persulfate was analyzed. Three replicate experiments were designed for each group, with each conical flask containing one slow-release agent.

2.4.2. Batch Static Degradation Experiments

A series of experiments was formulated to investigate the impacts of the initial contaminant concentration, catalyst dosage, and the chemical properties of groundwater on pollutant degradation efficiency. These experiments were carried out at room temperature (25 ± 1 °C) in 125 mL amber conical flasks to inhibit photolytic reactions. Three replicates were performed for each experimental condition.

Catalyst dosage experiment: To examine the influence of catalyst dosage, solutions of SMZ and CIP at concentrations of 1 mg L⁻¹ were prepared. Based on the pKa values of CIP and SMZ and the zeta potential of Fe@N-BC, the pH of these solutions was adjusted to 3 with 100 mM HCl and NaOH. 100 mL of antibiotic solution, varying amounts of Fe@N-BC catalyst (0 g, 0.01 g, 0.02 g, 0.03 g, 0.04 g), and one double-coated slow-release material were added to each reaction system. The degradation reactions were performed in a constant-temperature shaker at 25 °C and 150 rpm. At specified intervals, 1 mL of each reaction mixture was withdrawn into a 10 mL centrifuge tube, and 1 mL of methanol was added to terminate the reaction. The tube was subsequently centrifuged, and the supernatant was filtered through a 0.22 µm filter before undergoing UHPLC analysis to measure the concentrations of CIP and SMZ.

Initial pollutant concentration experiment: SMZ and CIP solutions with initial antibiotic concentrations of 1 mg L⁻¹, 5 mg L⁻¹, and 10 mg L⁻¹ were prepared and adjusted to pH 3. Then, 0.03 g Fe@N-BC and the premade optimized double-coated slow-release material were added to each reaction system. The reactions were performed in a constant-temperature shaker at 150 rpm. At specified intervals, 1 mL of each reaction mixture was withdrawn into a 10 mL centrifuge tube, and 1 mL of methanol was added to terminate the reaction. The tube was centrifuged, and the resulting supernatant was passed through a 0.22 µm filter before UHPLC analysis to determine the concentrations of CIP and SMZ.

Effect of groundwater chemical characteristics: Explore the effects of different pH values (pH 3, 5, 7, 9), humic acid concentrations (10 mg L⁻¹, 50 mg L⁻¹, 100 mg L⁻¹), and the presence of other anions (Cl⁻, PO₄³⁻, NO₃⁻, HCO₃⁻) on pollutant degradation rates. While repeating the above degradation experiment steps, control the above variables to observe the degradation of pollutants.

2.4.3. Identification of Reactive Species

To study the mechanism by which the double-coated slow-release materials degrade SMZ and CIP and to explore the dominant radicals activated by persulfate, tert-butyl alcohol (TBA), ethanol (MeOH), and furfuryl alcohol (FFA) were utilized as quenching agents. TBA was used to quench •OH radicals, MeOH was chosen to quench both SO₄^•− and •OH radicals, and FFA was used to quench ¹O₂. To perform the radical quenching experiments, 1 mg L⁻¹ SMZ and CIP solutions were prepared. In total, 100 mL of antibiotic solution and 0.03 g of Fe@N-BC were added to conical flasks. Then, the pH of each flask was adjusted to 3, followed by the addition of 100 mM of the required quenching agent. Finally, a double-coated persulfate slow-release material was added to each system. The experiment was conducted over a 12 h period to investigate the impact of Fe@N-BC on antibiotic remediation in the presence of quenching agents.

2.4.4. Dynamic Sand Column Release Experiment

This experiment was performed utilizing an organic glass column (Figure 1a) that measured 30 cm in height and had an inner diameter of 5 cm. Screens were installed at both ends of the column to prevent clogging. The column was first filled with 23.5 cm quartz sand of varying particle sizes. This was followed by the addition of four blocks (each 1.5 cm high) of slow-release material to form a single slow-release layer. Finally, the column was then topped off with an additional 5 cm of quartz sand of varying particle sizes, which was compacted as it was added. Water was pumped upward from the bottom through the quartz sand column, with the flow rate regulated by a peristaltic pump. Periodic sampling was performed for analysis. The different experimental groups are detailed in Table S2.

2.4.5. Dynamic Sand Column Degradation Experiment

To perform this experiment, the glass column described in Section 2.4.4 was filled with quartz sand, four blocks of slow-release material (each 1.5 cm high), and 15 cm of the catalyst Fe@N-BC (Figure 1b). The effects of groundwater flow rate (0.5 mL min⁻¹, 1 mL min⁻¹, 2 mL min⁻¹), aquifer medium grain size (quartz sand mesh 40–60, 120–140, and 200), and initial antibiotic concentration (1 mg/L, 10 mg/L, 50 mg/L) on antibiotic degradation performance under dynamic conditions were investigated. Water was pumped from the bottom upward through the quartz sand column, and the water flow rate was controlled with a peristaltic pump. The water contained specified concentrations of SMZ or CIP, and periodic sampling was performed. The detailed parameters for the dynamic degradation experiment are provided in Table S3.

Figure 1. Sand column experimental device. ((a) Release Experimental device; (b) Degradation Experimental device).

Figure 1. Sand column experimental device. ((a) Release Experimental device; (b) Degradation Experimental device).

2.5. Determination of Persulfate, Antibiotics, and Degradation Intermediates

Persulfate content was determined using UV–visible spectrophotometry. In total, 0.1 mL of sample, 0.05 g of NaHCO₃, and 1 g of KI were added to a colorimeter tube, and the reaction was carried out for about 15 min after dilution to 10 mL. Absorbance was determined using a visible spectrophotometer set to a wavelength of 352 nm. Then, the standard curve was plotted. Finally, the absorbance at the sampling point was measured, and the corresponding persulfate concentration was calculated.

The concentrations of CIP and SMZ were detected using UPLC. Analysis was performed with a BEH C18 column (1.7 µm, 2.1 × 100 mm), and CIP and SMZ were measured using DAD detectors at wavelengths of 264 and 270 nm separately. The degradation intermediates were analyzed and identified using high-resolution liquid chromatography-mass spectrometry (Ultimate 3000 UHPLC-Q Exactive, ThermoFisher Scientific, Waltham, MA, USA). Additional details on the analysis and the specific measurement conditions are presented in the supporting information (SI Text 1).

2.6. Material Characterization

The surface morphology of the slow-release materials was observed with scanning electron microscopy (SEM, Hitachi SU3500, Tokyo, Japan) at different magnifications. The catalyst structure was analyzed using X-ray diffractometry (XRD, Bruker D8 Advance, Saarbrücken, Germany). Metal valence states and surface elemental compositions were determined using X-ray photoelectron spectroscopy (XPS, Thermo Scientific K-Alpha, USA). The surface area, pore size, and pore volume of the slow-release materials were determined by N2 adsorption (Micromeritics ASAP 2460, Norcross, GA, USA). The relevant characterization results of the material are shown in Supplementary Material Figures S3–S5.

3. Results and Discussion

3.1. Optimal Formulation of Double-Coated Persulfate Slow-Release Materials

3.1.1. Persulfate Release Performance of Inner Layer Materials with Different Formulations

Five groups of cubic slow-release materials with a side length of 1.3 cm were prepared with different component ratios to study the release characteristics and performance of single-layer persulfate slow-release materials. The specific ratios of these groups (denoted A₁ to A₅) are shown in Figure 2a, while Figure 2b displays the cumulative release rates of persulfate by the different single-layer materials. After 21 days of release, the cumulative release rates of persulfate for groups A₁, A₂, and A₃ were 55.41%, 66.76%, and 83.13%, in that order. This is mainly because the increase in silica sand will lead to an increase in the pore space of the material. As the material pores become larger, the contact area with the water flow will also increase, resulting in an increased release of persulfate. At the same time, with the increase in the release of persulfate, the corrosion rate of the material will accelerate, and the surface will form larger pores. For the same mass of persulfate, the cumulative release rate increased with the increasing ratio of silica sand to PCL. On the 21st day, groups A₁, A₄ and A₂, A₅ had cumulative persulfate release rates of 55.41%, 72.19% and 66.76%, 82.21%, respectively. This indicates that the cumulative persulfate release rate increased with increasing persulfate mass under a fixed ratio of silica sand to PCL. Thus, the release efficiencies of these single-layer materials can be regulated by modifying the ratio of silica sand to PCL and adjusting the quantity of persulfate. As shown in Figure 2b, the persulfate release rate from the single-layer materials gradually increased over time. The release pattern indicates that a relatively high release rate was observed during the initial stage of the process. However, the persulfate release rate gradually decreased with increasing release time. This uneven release rate is a common drawback encountered in the study of uniform single-type slow-release materials [37]. Therefore, subsequent experiments focused on creating and testing double-coated persulfate slow-release materials to address this deficiency.

3.1.2. Release Characteristics of Double-Coated Slow-Release Materials with Different Outer-Layer Material Ratios

Based on the established persulfate/silica sand/PCL inner layer mass ratio of 1:3:1.5, six groups (denoted B₁ to B₆) of double-coated slow-release materials were prepared with different outer-layer material ratios, and the release characteristics and effectiveness of these slow-release materials were evaluated [38]. Each inner layer had a side length of 1.3 cm, and these cubes were covered with varying outer-layer ratios. The overall double-coated materials were cubes with a side length of 1.55 cm, and the thickness of the outer layer was 0.25 cm. After 21 days of release, the cumulative persulfate release rates from groups B₁, B₂, B₃, and B₄ were 0.98%, 7.48%, 35.52%, and 47.27%, respectively. In contrast, after just 11 days, the cumulative release rates of persulfate for groups B₅ and B₆ reached 48.09% and 54.88% (Figure 2c). This demonstrates that the composition of the outer-layer material significantly impacts the release performance of these slow-release materials. When the outer-layer thickness was 0.25 cm and the outer-layer silica sand/PCL ratio was greater than or equal to 5, the release rate of the double-coated persulfate slow-release material was somewhat controlled compared to the single-layer slow-release material. However, an excessive release rate was still observed in the early stages. Conversely, when the silica sand/PCL ratio was less than or equal to 2, the outer layer was significantly less permeable, which resulted in the excessively slow release of persulfate. When the silica sand/PCL ratio was between 2 and 5 (2 < silica sand/PCL < 5), the outer layer maintained good permeability while effectively controlling the release rate of persulfate (Figure 2c). These results demonstrate that the impact of the outer-layer material composition on release performance should be thoroughly investigated when designing double-coated slow-release materials.

3.1.3. Release Characteristics of Double-Coated Slow-Release Materials with Different Outer-Layer Thicknesses

The three types of double-layered slow-release materials with an outer layer of silica sand/PCL ratio of 4:1 and thicknesses of 0.35 cm, 0.25 cm, and 0.15 cm were designated as C₁, B₄, and C₃. Three additional groups of double-coated slow-release materials with an outer-layer silica sand to PCL ratio of 3:1 and thicknesses of 0.35 cm, 0.25 cm, and 0.15 cm were denoted as C₂, B₃, and C₄, in that order. The cumulative persulfate release rate after 21 days for group C₁ was 40.55%, for group B₄ it was 47.27%, and for group C3, it was 58.93%, following the order of C₁, B₄, and C₃. Meanwhile, the cumulative persulfate release rates of the C₂, B₃, and C₄ groups after 21 days were only 2.36%, 35.52%, and 23.47%, respectively. These results demonstrate that under fixed outer-layer permeability, the dissolution–diffusion pathway of persulfate increases with increasing outer-layer thickness, which leads to a slower persulfate release rate [39].

Figure 2. Mass ratios of slow-release materials with different material ratios and outer thicknesses (a), release rates of single-layer slow-release materials with different composition ratios (b), release rates of double-coated slow-release materials with different outer-layer ratios (c), and release rates of double-coated slow-release materials with different outer thicknesses (d).

Figure 2. Mass ratios of slow-release materials with different material ratios and outer thicknesses (a), release rates of single-layer slow-release materials with different composition ratios (b), release rates of double-coated slow-release materials with different outer-layer ratios (c), and release rates of double-coated slow-release materials with different outer thicknesses (d).

In summary, during the design process of double-coated slow-release materials, the release rate of persulfate from the inner layer can be regulated by altering the outer-layer material composition and thickness. Taking into account the relationship between release time and release rate, subsequent experiments were performed using the slow-release material B₄ (inner-layer sodium persulfate/silica sand/PCL material ratio = 1:3:1.5, outer-layer silica sand/PCL material ratio = 4:1, thickness = 0.25 cm).

3.2. Degradation of SMZ and CIP in Water by Double-Coated Persulfate Slow-Release Materials

The slow-release oxidative properties of the double-coated persulfate slow-release material for organic pollutant decomposition in aqueous solution were illustrated using SMZ and CIP as target pollutants and Fe@N-BC as an activator. Based on the static release and the activator test results (Supplemetary Materials Texts S2 and S3), the B4 slow-release material was used in these degradation experiments.

Due to the gradual release and increase of persulfate concentration over time, there is a significant difference in concentration between before and after, and the degradation reaction is essentially a chemical reaction involving ·OH and ·SO₄⁻ generated by activation. Therefore, quasi-first-order kinetic equations (Equation (1)) and quasi-second-order kinetic equations (Equation (2)) are used to study the degradation reaction of Fe@N-BC. The performance of activated double-coated persulfate slow-release materials in degrading CIP and SMZ was evaluated. The fitting results are shown in Figure S6 of the Supplementary Information.

$$-ln{\displaystyle \frac{c_t}{c_0}}=k_{1}t$$

(1)

$${\displaystyle \frac{1}{c_t}}-{\displaystyle \frac{1}{c_0}}=k_{2}t$$

(2)

In the formula, C₀ and C_t are the initial antibiotic concentration and the antibiotic concentration at time t (mg/L), respectively; T is the reaction time (h); k₁ and k₂ are the reaction rate constants for the pseudo-first-order (h⁻¹) degradation kinetics and pseudo-second-order ((mg/L)⁻¹· h⁻¹) degradation kinetics equations, respectively.

3.2.1. Effect of Catalyst Concentration on Pollutant Degradation

The degradation of SMZ and CIP was evaluated under different Fe@N-BC catalyst concentrations, as shown in Figure 3a,b. Catalyst concentrations of 0.3 g L⁻¹ and 0.4 g L⁻¹ resulted in the complete degradation of SMZ within 24 h, while the systems with catalyst concentrations of 0.1 g L⁻¹ and 0.2 g L⁻¹ required 48 h for complete SMZ degradation. A comparable trend was noted in the degradation of CIP: complete CIP degradation occurred within 2 h in the presence of 0.3 g L⁻¹ and 0.4 g L⁻¹ catalyst, while 12 h was required to completely degrade the CIP with 0.1 g L⁻¹ and 0.2 g L⁻¹ catalyst. As the catalyst concentration increased from 0 g/L to 0.4 g/L, the degradation reaction rate constants of CIP increased from k₁, CIP = 0.01807 h⁻¹ to k₁, CIP = 2.8746 h⁻¹, and the reaction rate constants of SMZ increased from k₁, SMZ = 0.0074 h⁻¹ to k₁, SMZ = 0.38179 h⁻¹. Although the concentrations of SMZ and CIP in the control group without the catalyst also showed a decreasing trend, the degradation rates were below 30%. Thus, the degradation performance of the control group was not sufficient. These results demonstrate that higher catalyst concentrations led to higher SMZ and CIP degradation rates. This could be due to higher catalyst concentrations generating a large number of radicals at a higher rate, which would increase the chance of pollutants (SMZ and CIP) interacting with these radicals. Thus, the degradation rate of the pollutants was enhanced [40]. This demonstrates the effectiveness of the catalytic oxidation system. Based on the above experimental results, the catalyst dosage for subsequent experiments was chosen to be 0.3g/L. Additionally, the experimental results also indicate that under the same conditions, CIP was removed more efficiently than SMZ. This is because CIP is composed of a piperazine and a quinolone ring connected by a short chain, while the SMZ molecule consists of a benzene ring and an isoxazole ring linked by a long chain. Compared to SMZ, CIP has a higher electron density, making it more susceptible to attack by SO₄•⁻ and •OH. Thus, a higher CIP removal rate can be achieved [41].

3.2.2. Effect of Initial SMZ and CIP Concentration on Pollutant Degradation

The degradation performance of the double-coated persulfate slow-release material was evaluated with different initial concentrations of SMZ and CIP, as shown in Figure 3c,d. After the initial concentration increases, the degradation rate of SMZ decreased. After 12 h of reaction, the degradation rate decreased from 100% (1 mg L⁻¹) to 50.18% (5 mg L⁻¹) and 23.2% (10 mg L⁻¹). In contrast, the initial concentration of CIP can degrade to 100% at 1–10 mg L⁻¹, but the time to achieve complete degradation varies from 2 h (mg L⁻¹) to 4 h (mg L⁻¹) and 8 h (mg L⁻¹). These results confirm that higher initial concentrations of SMZ and CIP lead to lower degradation rates within the same time period. This is because the free radicals generated by the activation of S₂O₈²⁻ released from the double-coated persulfate slow-release material were sufficient for the removal of low SMZ/CIP concentrations from the system [42]. However, at higher SMZ/CIP concentrations, the release of S₂O₈²⁻ was not sufficient to completely oxidize and degrade the SMZ/CIP, leading to reduced degradation rates. To avoid wasting persulfate or experiencing insufficiently low degradation rates, the ratio or dosage of the double-coated slow-release material should be chosen based on the pollutant concentrations at the actual application sites.

3.2.3. Effect of Initial pH on Pollutant Degradation

The degradation of SMZ and CIP by the double-coated persulfate slow-release material was investigated at initial pH values of 3, 5, 7, and 9, as shown in Figure 3e,f. Within the same time, the SMZ and CIP degradation rates both decreased with increasing pH. It took 12 h to reach 100% degradation of SMZ at initial pH 3, while at pH 5, the time required was 36 h. However, at pH 7 and 9, complete degradation was not achieved even after 48 h. When pH = 5, CIP can achieve complete degradation within 24 h, while when pH = 3, the time for CIP to reach complete degradation is shortened to 2 h. At pH 7 and after 24 h, CIP degradation was 97% and 91% at pH 10. For both SMZ and CIP, the time required for complete degradation increased as the initial pH increased. This phenomenon is related to the oxidizing properties of persulfate (S₂O₈²⁻) under different pH conditions. Under acidic conditions, S₂O₈²⁻ can react with H⁺ to form sulfate radicals (SO₄^−·). These sulfate radicals will subsequently react with H₂O to form ·OH (Equations (3)–(5)), thus promoting the activation of persulfate [43]. Under alkaline conditions, SO₄^−· reacts more readily with OH⁻ to form ·OH (Equations (6) and (7)), which has a lower oxidation potential than SO₄^−· [44]. In addition, the decrease in degradation performance may be due to the fact that iron-doped catalysts are the main active sites for PDS activation, and the precipitation of Fe (OH)₃ at alkaline pH inhibits the production of SO₄⁻ radicals (Equations (8) and (9)) [44,45].

$$S_2{O}_8^{2-}+H^{+}\to {HS}_2{O}_8^{-}$$

(3)

$${HS}_2{O}_8^{-}\to {SO}_4^{-.}+{SO}_4^{2-}+H^{+}$$

(4)

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

(5)

$$S_2{O}_8^{2-}+O{H}^{-}\to S{O}_4^{2-}+S{O}_4^{-\bullet }+H_{2}O$$

(6)

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

(7)

$${Fe}^{2+}+{S_2{O}_8}^{2-}\to {Fe}^{3+}+{SO}_4^{-}·+{SO}_4^{2-}$$

(8)

$${Fe}^{3+}+3O{H}^{-}\to {Fe\left(OH\right)}_3\downarrow$$

(9)

3.2.4. Effect of Humic Acid Concentration on Pollutant Degradation

The presence of humic acid had a substantial effect on the degradation of SMZ and CIP, as shown in Figure 3g,h. SMZ degradation decreased in the presence of humic acid. Degradation decreased from 100% (Control, 0 mg L⁻¹ humic acid) to 97.25% (10 mg L⁻¹), 77.01% (50 mg L⁻¹) and 62.75% (100 mg L⁻¹). In the CIP system, CIP was completely degraded within 24 h under humic acid concentrations of 50 mg L⁻¹ or lower, but the CIP removal rate decreased to 93.67% when the humic acid concentration was increased to 100 mg L⁻¹. In general, higher concentrations of humic acid resulted in lower degradation rates, indicating that humic acid had a strong inhibitory effect on the degradation of SMZ and CIP by the persulfate slow-release material. This inhibitory effect occurs because humic acid competes with the pollutants (SMZ and CIP) for the reactive radicals in the degradation system. Consequently, fewer interactions between the pollutants and the radicals occur, which hinders the oxidative degradation of the pollutants [46,47,48]. Therefore, when remediating actual contaminated groundwater sites, the impact of organic components in the groundwater on the pollutant degradation efficiency should be considered. Increasing the persulfate slow-release material dosage is expected to mitigate this effect to some extent.

Figure 3. Effects of catalyst dosing (a,b), initial pollutant concentration (c,d), initial pH (e,f), and humic acid dosing (experimental conditions: pH = 3, initial pollutant concentration 1 mg/L, catalyst 0.3 g/L, a new double-coated slow-release material, temperature 25 °C) (g,h) on SMZ and CIP degradation.

Figure 3. Effects of catalyst dosing (a,b), initial pollutant concentration (c,d), initial pH (e,f), and humic acid dosing (experimental conditions: pH = 3, initial pollutant concentration 1 mg/L, catalyst 0.3 g/L, a new double-coated slow-release material, temperature 25 °C) (g,h) on SMZ and CIP degradation.

3.2.5. Effect of Anion Concentration on Pollutant Degradation

The impact of various anions at different concentrations on the degradation of SMZ and CIP by the double-coated persulfate slow-release material was evaluated, as shown in Figure 4. In systems containing 10 mmol L⁻¹, 50 mmol L⁻¹, and 100 mmol L⁻¹ of Cl⁻, SMZ was completely degraded after 24 h, 36 h, and 48 h, respectively, while CIP was completely degraded after 4 h, 6 h, and 24 h, respectively. Thus, as the Cl⁻ concentration increased, the degradation rates of both SMZ and CIP decreased. The inhibitory effect of Cl⁻ on the degradation of SMZ and CIP by persulfate increased with increasing Cl⁻ concentration, and Cl⁻ exhibited a stronger inhibitory effect on the degradation of SMZ degradation compared to that of CIP. This is attributed to the formation of low-reactive chlorine radicals (Cl·) through secondary reactions between Cl⁻ in the solution and ·OH and SO₄⁻· in the oxidation system (Equations (10) and (11)) [49]. The reduction potential of Cl· is 2.41 V, equivalent to SO₄⁻· radicals, which is lower than that of SO₄⁻·. Thus, the generation of Cl· slightly reduces the SMZ and CIP degradation efficiency of persulfate.

$$S{O_4}^{-}·+C{l}^{-}\leftrightarrow S{O_4}^{2-}+Cl·$$

(10)

$$·OH+C{l}^{-}\to O{H}^{-}+Cl·$$

(11)

The addition of NO₃⁻, PO₄³⁻, and HCO₃⁻ significantly inhibited the degradation of SMZ and CIP (Figure 4), and the SMZ and CIP degradation rates noticeably declined with increasing concentrations of all three anions. In a system containing 10 mmol L⁻¹ NO₃⁻, the complete degradation time of SMZ is 36 h, while when the concentration of NO₃⁻ reaches 50 mmol L⁻¹, the required time is 48 h, while in the system with 100 mmol L⁻¹ NO₃⁻, the SMZ degradation rate was less than 95% after 48 h. For CIP, the systems with 10 mmol L⁻¹, 50 mmol L⁻¹, and 100 mmol L⁻¹ NO₃⁻ showed CIP degradation rates of 100% after 24 h, 36 h, and 48 h of reaction, respectively. This suggests that NO₃⁻ slightly inhibited the degradation of SMZ and CIP, and that the inhibition of CIP degradation was relatively more pronounced. The slight inhibition by NO₃⁻ addition may be due to the fact that NO₃⁻ can react with SO₄⁻· to form NO₃$·$ (Equation (12)), which may result in the scavenging of sulfate radicals to a certain extent. However, the less significant inhibitory effect of NO₃⁻ may be due to the lower reaction rate constants of NO₃⁻ with HO• and SO₄•⁻ radicals, i.e., <1 × 10⁵ M⁻¹s⁻¹ [50] and 2.1 × 100 M⁻¹s⁻¹ [51]>, and some people believe that adding NO₃⁻ may enhance the adsorption of pollutants by biochar materials [52], which may be the reason why the inhibitory effect of NO₃⁻ is not as significant as that of HCO₃⁻ and PO₄³⁻.

$${SO}_4^{-\bullet }+N{O_3}^{-}\to S{O_4}^{2-}+N{O}_3$$

(12)

The presence of PO₄³⁻ exhibited a stronger inhibitory effect on the degradation of SMZ and CIP than that of NO₃⁻, and PO₄³⁻ also inhibited the degradation of SMZ more than that of CIP. In systems with 10 mmol L⁻¹, 50 mmol L⁻¹, and 100 mmol L⁻¹ PO₄³⁻, SMZ degradation rates of less than 80% were achieved after 24 h, and after 48 h, the degradation rates were still below 90%. For CIP, degradation rates below 90% were achieved after 24 h in the presence of PO₄³⁻. A CIP degradation of 100% was achieved in the system containing 10 mmol L⁻¹ PO₄³⁻ after 36 h, and in the presence of 50 mmol L⁻¹ PO₄³⁻, the CIP degradation rate reached 100% after 48 h.

Figure 4. Effect of anions ((a,b), added Cl⁻; (c,d), added NO₃⁻; (e,f), added PO₄³⁻; (g,h), added HCO₃⁻) on SMZ and CIP degradation (experimental conditions: pH = 3, initial pollutant concentration 1 mg/L, catalyst 0.3 g/L, a new double-coated slow-release material, temperature 25 °C).

Figure 4. Effect of anions ((a,b), added Cl⁻; (c,d), added NO₃⁻; (e,f), added PO₄³⁻; (g,h), added HCO₃⁻) on SMZ and CIP degradation (experimental conditions: pH = 3, initial pollutant concentration 1 mg/L, catalyst 0.3 g/L, a new double-coated slow-release material, temperature 25 °C).

HCO₃⁻ exhibited an even stronger inhibitory effect on SMZ and CIP degradation compared to the other evaluated anions, and the effect of HCO₃⁻ on CIP degradation was stronger than that on SMZ degradation [53]. The second-order rate constant for the reaction of HCO₃⁻ with SO₄⁻ is 6.0 × 10⁶ M⁻¹s⁻¹, and the reaction rate constant with ·OH is 8.5 × 10⁶ M⁻¹s⁻¹ [54]. After 24 h, CIP degradation rates below 73% and SMZ degradation rates below 75% were achieved in systems containing different HCO₃⁻ concentrations. The inhibitory effect of the evaluated anions on the degradation of SMZ and CIP follows the order: Cl⁻ < NO₃⁻ < PO₄³⁻ < HCO₃⁻. The anions PO₄³⁻ and HCO₃⁻ strongly inhibit the degradation of SMZ and CIP due to their ability to scavenge the SO₄⁻· and ·OH radicals generated during the reaction (Equations (13)–(18)). The reaction rate constants of PO₄³⁻ with ·OH and SO₄⁻· are 1.1 × 10⁹ M⁻¹s⁻¹ [54] and 1.2 × 10⁶ M⁻¹s⁻¹ [55]. This reduces the number of radicals available to react with SMZ and CIP, which inhibits their degradation [56,57].

$$·OH+HC{O_3}^{-}\to O{H}^{-}+HC{O}_3·$$

(13)

$$·OH+C{O_3}^{2-}\to O{H}^{-}+C{O_3}^{-}·$$

(14)

$${SO}_4^{-\bullet }+HC{O_3}^{-}\to S{O_4}^{2-}+HC{O}_3·$$

(15)

$${SO}_4^{-\bullet }+C{{O_3}^2}^{-}\to S{O_4}^{2-}+C{O_3}^{-}·$$

(16)

$${SO}_4^{-\bullet }+{PO}_4^{3-}\to {SO}_4^{2-}+{PO}_4^{2-}·$$

(17)

$$·OH+{PO}_4^{3-}\to {OH}^{-}+{PO}_4^{2-}·$$

(18)

3.3. Release Characteristics of Persulfate Slow-Release Materials in One-Dimensional Columns

3.3.1. Influence of Flow Rate on Release Characteristics of Persulfate Slow-Release Material in One-Dimensional Columns

The release performance of the double-coated persulfate slow-release material in a one-dimensional column was studied under different water flow rates over a 60-day monitoring period, as shown in Figure 5a. The specific parameters of the dynamic repair experiment are detailed in Table S4. Under the evaluated flow rates, high effluent persulfate concentrations were observed within the first 10 days, followed by a gradual decrease and stabilization. This was consistent with the static release experiment results. As the release of persulfate increased, pores gradually appeared on the surface of the release material, and the decrease in residual persulfate caused the leaching of active components to stabilize [58]. The lowest effluent persulfate concentration was attained at a flow rate of 2 mL min⁻¹, while the highest effluent persulfate concentration was achieved at a flow rate of 0.5 mL min⁻¹ (Figure 5a). This is potentially because the high flow rate of 2 mL min⁻¹ results in insufficient contact time between the water and the release material, thus reducing the leaching of persulfate. Conversely, at a low flow rate of 0.5 mL min⁻¹, adequate contact and sufficient flushing force are established. Furthermore, the release period of persulfate decreased with an increasing flow rate. At a flow rate of 0.5 mL min⁻¹, continuous and effective persulfate release was maintained throughout the monitoring period. Conversely, at a flow rate of 2 mL min⁻¹, the persulfate release period was shorter. The effective persulfate release times follow the order 0.5 mL min⁻¹ > 1 mL min⁻¹ > 2 mL min⁻¹, indicating that an increase in flow rate shortens the lifespan of the double-coated slow-release material. Considering the time constraints of the experiment and in order to be able to show more obvious differences in the experimental results, 1 mL/min was subsequently chosen for the experiment.

3.3.2. Influence of Medium Permeability on the Release Characteristics of Persulfate Slow-Release Material in One-Dimensional Columns

One-dimensional columns were filled with quartz sand with different particle sizes (40–60 mesh, 120–140 mesh, and 200 mesh) to evaluate the influence of permeability on persulfate release by the double-coated slow-release material, as shown in Figure 5b. This experiment was performed with a water flow rate of 1 mL min⁻¹. As the mesh size of quartz sand increases, the outflow concentration of persulfate also increases, from 61.48 mg L⁻¹ (40–60 mesh) to 199.73 mg L⁻¹ (120–140 mesh) and 189.21 mg L⁻¹ (200 mesh). Irrespective of the permeability conditions, the persulfate concentration exhibited an initial increase followed by a gradual decline, eventually stabilizing between 2 mg L⁻¹ and 27 mg L⁻¹. Within the first 10 days, the released concentrations of persulfate increased with increasing quartz sand. However, as the monitoring time progressed, the differences gradually decreased and stabilized. The sand column filled with 200 mesh quartz sand achieved complete release of persulfate on the 54th day, while the 120–140 mesh sand column showed this situation on the 55th day. Within the same release time, the persulfate effluent concentrations of the columns with different quartz sand particle sizes did not differ significantly. This suggests that within the experimental range, the release characteristics of the buffering material are only minimally influenced by the permeability coefficient of the quartz sand. The initial concentration differences are mainly due to the larger particle size of the quartz sand, which results in higher permeability and diffusion coefficients. Consequently, the average flow velocity of the water is higher, leading to a decrease in persulfate concentration with increasing quartz sand particle size. Additionally, the larger amount of persulfate encapsulated within the slow-release material in the initial stages resulted in a slightly higher concentration of persulfate released through dissolution–diffusion mechanisms compared to later stages [59]. With increasing persulfate release, pores gradually appeared in the surface layer of the slow-release material, and the remaining amount of persulfate decreased, leading to a stabilization in the release of active components [60].

Figure 5. Release of persulfate under different water flow rates (a) and different permeability coefficients (b).

Figure 5. Release of persulfate under different water flow rates (a) and different permeability coefficients (b).

3.4. Oxidative Degradation of SMZ and CIP by Double-Coated Persulfate Slow-Release Material in One-Dimensional Sand Columns

3.4.1. Influence of Groundwater Flow Velocity on SMZ and CIP Degradation

The effectiveness of the double-coated persulfate slow-release material for the remediation of SMZ and CIP in one-dimensional sand columns was evaluated for 120 h under different groundwater flow rates, as shown in Figure 6a,b. After the effluent concentrations stabilized, SMZ removal rates of 69.5%, 65.4%, and 61% were achieved under flow rates of 0.5 mL min⁻¹, 1 mL min⁻¹, and 2 mL min⁻¹, respectively, while the corresponding CIP removal rates were 86.4%, 82%, and 75.9%, respectively. Thus, the flow rate has a certain influence on the degradation of SMZ and CIP by the double-coated slow-release material, with better remediation effectiveness achieved at lower flow rates. Two main factors influence this trend. On the one hand, the flow rate affects the release of persulfate: higher flow rates result in lower persulfate effluent concentrations, which may not be sufficient for the remediation of SMZ and CIP. Thus, higher flow rates lead to lower remediation effectiveness, which is consistent with the dynamic release experiment results. On the other hand, at lower flow rates, the persulfate released by the double-coated slow-release material remains in the activation zone for a longer period. Consequently, the released persulfate has enough time to adequately activate the oxidation of pollutants [61,62,63]. Under similar conditions, the remediation effectiveness of the system for CIP is superior to that for SMZ, which is in agreement with the static remediation results.

Long-term dynamic remediation monitoring was conducted for the experimental group with a flow rate of 1 mL/min and an initial antibiotic concentration of 10 mg/L to study the long-term remediation capability of the double-coated persulfate slow-release material for CIP and SMZ in groundwater. The longest monitoring time was 34 days. As shown in Figure S7, the antibiotic concentration initially decreased to a minimum and then slowly rose and stabilized, gradually increasing to the initial inflow concentration. This change occurred in several stages. In the first stage, within the first 1 h of remediation, the double-coated persulfate slow-release material provided persulfate concentrations suitable for remediation, achieving optimal remediation effectiveness. In the second stage, within 240 h of remediation (i.e., before the released persulfate concentration decreased and stabilized), the antibiotic effluent concentration increased as remediation effectiveness rose and stabilized. In the third stage, the antibiotic effluent concentration continuously increased. This was partly due to the decreasing concentration of released persulfate, which was not sufficient for degrading the antibiotic, and partly due to the depletion of pores and active sites on the catalyst surface as the reaction progressed, which meant that the released persulfate was not able to activate the generated radicals. Consequently, the antibiotic removal efficiency rapidly decreased until the concentration of the effluent reached an equilibrium with that of the influent.

3.4.2. Effect of Quartz Sand Mesh on SMZ and CIP Degradation

The remediation effectiveness of the double-coated persulfate slow-release material under different permeability conditions (quartz sand particle sizes of 40–60 mesh, 120–140 mesh, and 200 mesh) is shown in Figure 6c,d. When the pollutant effluent concentration stabilized, SMZ degradation rates of 65.4%, 57.6%, and 56.2% were achieved in the columns with 40–60 mesh, 120–140 mesh, and 200 mesh quartz sand, respectively, while the corresponding CIP degradation rates were 82%, 76.8%, and 72.8%, respectively. The removal rates of both antibiotics initially increased and then decreased. This indicates that the permeability of the filling medium influences the CIP and SMZ remediation effectiveness of the double-coated persulfate slow-release material. As the permeability of the medium decreases, the pollutant remediation efficiency decreases, which is consistent with the dynamic release experimental results.

On the one hand, as the permeability coefficient decreases, the persulfate release rate decreases, leading to lower CIP and SMZ degradation rates due to the lower persulfate concentration in the system. On the other hand, lower permeability means that smaller gaps exist between the quartz sand particles, leading to an increase in the relative velocity of the contaminated groundwater. This results in a decreased contact time between the pollutants and the active reaction zone, resulting in insufficient reaction between the pollutants and persulfate. Consequently, the remediation efficiency declines. Therefore, in practical applications, the dosage or composition of slow-release materials should be reasonably designed according to the permeability of the contaminated site strata to achieve effective remediation of pollutants.

Figure 6. Effects of flow rate (a,b), permeability coefficient (c,d), and initial pollutant concentration (e,f) on the degradation rates of SMZ and CIP.

Figure 6. Effects of flow rate (a,b), permeability coefficient (c,d), and initial pollutant concentration (e,f) on the degradation rates of SMZ and CIP.

3.4.3. Effect of Initial Antibiotic Concentration on SMZ and CIP Degradation

The CIP and SMZ remediation effectiveness of the double-coated persulfate slow-release material at initial pollutant concentrations of 1 mg L⁻¹, 10 mg L⁻¹, and 50 mg L⁻¹ is shown in Figure 6e,f. When the effluent concentration stabilized, SMZ removal rates of 100%, 65.4%, and 43.06% were achieved at initial pollutant concentrations of 1 mg L⁻¹, 10 mg L⁻¹, and 50 mg L⁻¹, respectively, while the corresponding removal rates of CIP were 100%, 82%, and 59.06%, respectively. As the pollutant concentration increased, remediation effectiveness decreased. This is because only a limited amount of persulfate is released by the double-coated slow-release material within a certain period of time, and the active substances generated by Fe@N-BC activation are not sufficient for the oxidation and degradation of high concentrations of pollutants. Under low pollutant concentrations, the amount of persulfate released by the double-coated slow-release material is sufficient to degrade pollutants [64]. In practical applications, the remediation of pollutants at different concentrations can be achieved by adjusting the dosage or the formulation of slow-release materials.

3.5. SMZ and CIP Degradation Mechanism in the Fe@N-BC-Activated Persulfate System

3.5.1. Identification of Active Species

The free radicals ·OH and SO₄⁻· are commonly present in activated persulfate systems, and ¹O₂ has also been reported to be an important active species [65,66,67]. In order to determine the main active substances in the degradation of CIP and SMZ by Fe@N-BC-activated persulfate, we performed quenching experiments using the quenching agents TBA, MeOH, and FFA. TBA reacts with ·OH and SO₄⁻· at rates of k₂ = 6 × 108 M⁻¹·S⁻¹ and k₂ = 4 × 10⁵ M⁻¹·S⁻¹, respectively [68]. The reaction rate of TBA with SO₄⁻· is much lower than that with ·OH. Thus, TBA was used as a quenching agent for ·OH. MeOH reacts with ·OH and SO₄⁻· at the rate of k₂ = 9.7 × 10⁸ M⁻¹·S⁻¹ and k2 = 2.5 × 10⁷ M⁻¹·S⁻¹, respectively [69]. The reaction rate of FFA with ¹O₂ is k₂ = 1.2 × 10⁸ M⁻¹·S⁻¹ [70]. Figure 7 shows the degradation of CIP and SMZ after the addition of 100 mM of each quenching agent to the reaction system.

As shown in Figure 7, when MeOH was added and reacted for 12 h, the degradation rate of CIP decreased to 27.31% and that of SMZ decreased to 19.32%. The degradation rate of the pollutants after the addition of TBA decreased but to a lesser extent than that of the group with the addition of MeOH (degradation rates after the addition of TBA: CIP (46.12%) and SMZ (37.46%)). This result indicated that SO₄⁻· and ·OH were present in the system. FFA reacted rapidly with ¹O₂ [71], so FFA was chosen as a quencher to verify the presence of ¹O₂. The removal rate of CIP decreased to 52.18% after the addition of FFA, while the removal rate of SMZ was 42.02%. This indicates that ¹O₂ is generated in the system and participates in the degradation process of pollutants.

Figure 7. Degradation of pollutants in the presence of quenching agents.

Figure 7. Degradation of pollutants in the presence of quenching agents.

3.5.2. SMZ and CIP Degradation Products and Pathways in the Fe@N-BC-Activated Persulfate System

The intermediates generated from the oxidative degradation of SMZ and CIP were analyzed using liquid chromatography–mass spectrometry (LC-MS). Based on the intensity of the molecular ion peaks in the TIC chromatograms and literature information, several possible degradation pathways of SMZ and CIP in the Fe@N-BC-PS system were proposed, as shown in Figure 8.

The proposed degradation pathways for SMZ are displayed in Figure 8a. First path 1 is the N-S bond in the pre-sulfamethoxazole is attacked and broken to generate P₁ (m/z 98) and P₂ (m/z 155) [72], the C-S bond in P₂ is further decomposed and replaced by hydroxyl group to generate P₃ (m/z 110), and the amino group of P₃ is further oxidized to generate P₄ (m/z 137) [73]. Pathway 2 is the hydrogenation of the isoxazole ring to produce P₅ (m/z 171) sulfonamide. Path 3 is the attack of the N-C bond of the isoxazole ring [74] to produce P₆ (m/z 227), which is then further degraded to P₇ (m/z 192). Path 4 is the appearance of hydroxyl group at the benzene ring and isoxazole ring opening and generates carbonyl group by substitution reaction and dehydration to produce P₈ (m/z 224), and then the amino group on the benzene ring is further oxidized to remove the hydroxyl group, and the isoxazole ring is further decomposed to produce P₉ (m/z 227) [75]. Pathway 5 is the substitution of the amino group with the hydroxyl group to produce the intermediate product P₁₀ (m/z 269). As oxidation continues, these intermediates are mineralized into CO₂, H₂O, and other small molecules.

The most common degradation pathways for CIP include oxidative degradation of the piperazine side chain, hydroxylation of the quinolone ring, and F-ion -OH substitution [76,77]. First, pathway 1, the piperazine epoxidation, opens the ring and produces the dialdehyde derivative A₁ (m/z 362), which then loses two of the formaldehyde and loses the secondary ammonia nitrogen to produce A₂ (m/z 291) [78]. Path 2 is also the process of opening the piperazine ring. First, the piperazine ring amino nitrogen acetylation followed by demethylation results in the formation of the intermediate product A₃ (m/z 360), then hydrogenation occurs on the formyl group to produce A₄ (m/z 361). Path 3 is a change on the quinolone ring, where -OH replaces -COOH on the quinolone ring [79] to produce B1 (m/z 302). Pathway 4 is also a change on the quinolone ring with cleavage of the cyclopropyl group to generate B₂ (m/z 291), followed by further defluorination, decarboxylation, and loss of the piperazine ring to generate B₃ (m/z 149). Pathway 5 is the substitution of -OH for -F to produce C₁ (m/z 330), and then the piperazine ring undergoes oxidation to produce C₂ (m/z 304). Path 6 is the opening of both the piperazine ring and the quinolone ring [80], generating D₁ (m/z 308), followed by further decomposition of the piperazine ring, further cleavage of the quinolone ring to generate D₂ (m/z 171), and finally complete detachment at the piperazine ring to generate D₃ (m/z 155). Through further degradation, the intermediates of CIP are fully mineralized into simple inorganic molecules, including NH₄⁺, F⁻, CO₂, and H₂O.

Figure 8. Proposed degradation pathways of SMZ (a) and CIP (b) in the Fe@N-BC-PS system.

Figure 8. Proposed degradation pathways of SMZ (a) and CIP (b) in the Fe@N-BC-PS system.

4. Conclusions

In summary, this work provides an effective strategy for developing slow-release oxide materials. Compared with other oxidants, persulfate oxidants have the characteristics of strong oxidation, good stability, wide pH range, and long free radical life [81,82,83,84]. The PCL selected for this material also to some extent solves the problem of easy solubility in organic pollutants of commonly used material paraffin, and the double-layer coating also to some extent solves the problem of rapid loss and poor later effect of single-layer materials [30,85,86,87]. In addition, the PCL/silica sand composite was capable of retaining a large fraction of the oxidative power of persulfate [88]. The degradation of SMZ and CIP in water was studied using Fe@N-BC to activate the persulfate released by the double-coated slow-release material. In this system, the degradation rate decreased with increasing pH and increasing initial pollutant concentration. In a one-dimensional column experiment performed to evaluate the dynamic degradation of SMZ and CIP using the double-coated slow-release material, an increase in the initial pollutant concentration, larger silica sand mesh size, and higher flow rate all led to a decline in degradation performance. Therefore, in the practical application of slow-release materials, a detailed site investigation should be conducted. The hydrogeological conditions and the type and concentration of pollutants need to be determined to specify the size, material ratio, and dosage of slow-release materials for effective environmental remediation. In the investigation of factors affecting the release of double-coated slow-release materials and the degradation of typical antibiotics in groundwater, the influence of groundwater temperature on the release of persulfate and the remediation of antibiotic pollution was not considered. In future research, the temperature of the reaction system can be controlled to investigate its effects on release and repair.