Enhanced Fenton Degradation of Tetracycline over Cerium-Doped MIL88-A/g-C₃N₄: Catalytic Performance and Mechanism

Abstract

Since enormous amounts of antibiotics are consumed daily by millions of patients all over the world, tons of pharmaceutical residuals reach aquatic bodies. Accordingly, our study adopted the Fenton catalytic degradation approach to conquer such detrimental pollutants. (Ce_0.33Fe) MIL-88A was fabricated by the hydrothermal method; then, it was supported on the surface of g-C₃N₄ sheets using the post-synthetic approach to yield a heterogeneous Fenton-like (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst for degrading the tetracycline hydrochloride drug. The physicochemical characteristics of the catalyst were analyzed using FT-IR, SEM-EDX, XRD, BET, SEM, and XPS. The pH level, the H₂O₂ concentration, the reaction temperature, the catalyst dose, and the initial TC concentration were all examined as influencing factors of TC degradation efficiency. Approximately 92.44% of the TC was degraded within 100 min under optimal conditions: pH = 7, catalyst dosage = 0.01 g, H₂O₂ concentration = 100 mg/L, temperature = 25 °C, and TC concentration = 50 mg/L. It is noteworthy that the practical outcomes revealed how the Fenton-like process and adsorption work together. The degradation data were well-inspected by first-order and second-order models to define the reaction rate. The synergistic interaction between the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ components produces a continuous redox cycle of two active metal species and the electron-rich source of g-C₃N₄. The quenching test demonstrates that ^•OH is the primary active species for degrading TC in the H₂O₂–(Ce_0.33Fe) MIL-88A/10%g-C₃N₄ system. The GC-MS spectrum elucidates the yielded intermediates from degrading the TC molecules.

1. Introduction

There is no doubt that water is a lifeline for everyone. Water contamination has progressed significantly in the last few decades with restricted water supplies, making researchers concerned about wastewater treatment. Water pollution describes the contamination of water bodies by potentially dangerous substances, such as industrial and agricultural waste effluents that contain heavy metals, aromatic compounds, dyes, pesticides, and pharmaceuticals [1]. Pharmaceutical pollutants have risen, especially during the COVID-19 pandemic, which has revealed unexpected threatening effects due to the rapid rise in the number of COVID-19 patients who need antibiotics. There has been a corresponding increase in the presence of these drugs, which are improperly disposed of through sewage or medical waste [2]. These pharmaceutical contaminants seriously impede human health and aquatic life to a frightening degree [3]. Consequently, researchers have presented efforts to conquer this inevitable danger [4].

Tetracycline hydrochloride (TC) is a broad-spectrum antibiotic utilized for treating bacterial infections, including acne, urinary tract infections, respiratory tract infections, and skin, stomach, and intestine infections. Notably, TC is present in high concentrations because it is not entirely digestible, and between 50 and 80 percent of the dosage is excreted in the urine [2]. Thus, TC residues end up in water bodies, increasing the chances of bacteria developing antibiotic resistance [5]. Thereby, to prevent this, a variety of remediation methods have been developed for the elimination of TC from wastewater, including membrane separation [6], microbial degradation [7], electro-oxidative degradation [8], photocatalysis [9], adsorption [10], and advanced oxidation processes (AOPs) [11]. Among these methods, AOPs have drawn particular attention from researchers due to their eco-friendliness and high efficiency and reproductivity in eliminating TC [12,13].

In 1894, Fenton disclosed the Fenton reaction, which aims to convert pollutants to less harmful substances [13]. The conventional homogeneous Fenton reaction system is a mixture of Fe²⁺ and H₂O₂ in highly acidic conditions to generate Fe³⁺ and hydroxyl radicals ^⦁OH with a high redox potential (E^o of ^⦁OH/H₂O₂ = 2.73 V) as the main reactive oxygen species (ROS) to degrade contaminants in a non-selective way [11,14]. Then, the reaction between Fe³⁺ and H₂O₂ reduces it to Fe²⁺, so ^⦁OH will be produced endlessly by a continuous redox cycle (Fe²⁺/Fe³⁺) [13,15]. The traditional Fenton reaction has flaws such as not being recyclable, a large consumption of H₂O₂, acidic pH, and iron leaching. Consequently, the insoluble solid catalyst-based heterogeneous Fenton-like reaction has been developed to enhance the issues mentioned above since it is inexpensive, has mild operating conditions, recovers quickly, and does not produce sludge [12,14,16,17]. The Fenton-like catalyst is an insoluble solid catalyst that imitates the behavior of the Fenton reaction. Furthermore, the Fenton-like catalyst can include other non-ferrous multivalent transition metals (Ni, Co, Ag, and Au), metal oxides (Mn₃O₄, NiS, and CeO₂), carbon-based materials (activated carbons, carbon nanotubes, graphite, graphene and reduced graphene oxide) as well as other new type materials (e.g., polyoxometalates and metal–organic frameworks, MXene, and layered double hydroxide) [18,19]. On the other hand, adsorption is a potent approach because of its distinctive benefits, including being simple, economical, energy-saving, and environmentally friendly, and its ease of regeneration and high efficiency [20]. According to pioneering studies, the Fenton process and adsorption work together to improve removal efficiency for organic contaminants [21].

Porous materials comprising metal–organic frameworks (MOFs) have exhibited great interest as super-eminent adsorbents for purifying organic contaminant-containing wastewater. Metal–organic frameworks are engineered from metal ions bound by organic ligands. Noteworthy, MOFs are advanced substances with promising features, such as water stability, high thermal and chemical stability, and functionalization simplicity. Also, MOFs have a large surface area, a variable pore size, and elastic interior surface features. These properties make them valuable in numerous applications, including gas storage, separation, sensors, solar cells, batteries, supercapacitors, drug delivery, membranes, adsorption, and catalysis [22,23]. Until now, the Fe-based MIL family, which comprises MIL-88A(Fe), MIL-88B(Fe), MIL-53(Fe), MIL-100(Fe), and MIL-101(Fe), has demonstrated significant promise as a successful Fenton-like catalyst [24,25]. Strikingly, MIL-88A(Fe) is the most widely used since it is easy to produce and has strong chemical and water stability [12]. Also, it is environmentally friendly because it can be prepared in an aqueous medium instead of a toxic solvent like dimethyl formamide [26]. Since MIL-88A(Fe) only includes Fe³⁺ species with a weak Fenton degradation aptitude, it is necessary to enhance the catalytic effectiveness of the pure MIL-88A(Fe) [13]. Adjusting the Lewis acidity of MIL-88A(Fe) through cerium (Ce) doping may result in more intrinsic ligand missing defects and improve the catalytic activity. Cerium is one of the most common rare-earth elements that is characterized by its rich redox characteristics, high electronic/ionic conductivity, adaptable coordination ability, and hard Lewis acid [27,28,29].

Carbon nitride is a 2D polymer with a layer-like structure that can be synthesized easily from cheap and abundant resources. g-C₃N₄ is a stable, low-cost, non-toxic, high thermal and chemical stability, and environmentally friendly material. Studies implied the favorability of using g-C₃N₄ as a Fenton-like catalyst during the degradation of organic pollutants. The g-C₃N₄ catalyst can activate hydrogen peroxide to generate hydroxyl radicals, which can then oxidize the organic contaminants. The activity, stability, accessibility of active sites, and surface area of g-C₃N₄ as a Fenton-like catalyst can be boosted by doping it with other elements such as iron or cobalt or by coupling with different materials like metal oxides, graphene, and MOF [30,31,32,33].

Herein, we report the synthesis of a novel heterogeneous Fenton-like catalyst for degrading the residual TC molecules in wastewater. The Fenton-like catalyst was fabricated by doping the Ce species onto the MIL-88A surface, and then the obtained Ce-doped MIL-88A was combined with g-C₃N₄. To optimize the best molar ratios between the catalyst components, a series of synthesis experiments were performed to fabricate Ce-doped MIL-88A/g-C₃N₄ with various molar ratios of the doped Ce species and g-C₃N₄. The as-fabricated (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst was fully characterized via XRD, XPS, SEM, SEM-EDX, BET, elemental mapping and FT-IR to investigate its structural, compositional, and morphological specifications. After several lab experiments, the best conditions for the Fenton-like reaction of TC by the H₂O₂–(Ce_0.33Fe) MIL-88A/10%g-C₃N₄ system were thoroughly determined. The potential degradation pathway of TC was supposed by the experimental results of the scavenging test and the XPS analysis. Furthermore, the stability of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst was assessed by performing a recycling test and defining the metal leaching concentration. The intermediate molecules during the Fenton-like degradation of TC by (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ were determined utilizing GC-MS.

2. Experimental

2.1. Materials

Ferric chloride hexahydrate 97% (FeCl₃·6H₂O) and Fumaric acid (C₄H₄O₄) 99.5% were acquired from Loba Chemie (Mumbai, India). Cerium oxalate, sodium hydroxide (97%, NaOH), hydrochloric acid (37%, HCl), and ethanol (99%, C₂H₅OH) were purchased from Merck, Rahway, NJ, USA. P-benzoquinone (PBQ) (99%) was supplied from Shanghai Chemical Reagent Co., Ltd. (Shanghai, China). Urea CO(NH₂)₂, Hydrogen peroxide (50%), and tetracycline hydrochloride (C₂₂H₂₅Cl N₂O₈, ≥ 99%, MW 480.9 g/mol) were provided by Sigma-Aldrich Co. (Darmstadt, Germany). Without any additional purification, all chemicals employed in the synthesis steps were used as obtained.

2.2. Synthesis of MIL-88A

The hydrothermal technique was utilized to manufacture MIL-88A [34]. Briefly, 10 mmol of FeCl₃.6H₂O and 10 mmol of Fumaric acid were dissolved into 50 mL ultrapure water and vigorously stirred using a magnetic stirrer for nearly 1 h at room temperature. The mixture was then poured into a 100 mL autoclave and heated in an oven for 24 h at 85 °C. After cooling down to ambient temperature without any external assistance, the obtained orange precipitate was collected using the centrifuge, purified using ethanol and deionized water for one time, and dried at 85 °C overnight.

2.3. Preparation of Ce-Doped MIL-88A

The same former procedure was applied to fabricate the Ce-doped MIL-88A(Fe) materials with an additional step, which included adding various amounts of cerium oxalate ranging from 5 to 50%. The Ce species was mixed to FeCl₃.6H₂O under stirring for 1 h, and then the rest of the preparation procedure of MIL-88A was completed as previously mentioned.

2.4. Synthesis of g-C₃N₃

To prepare the sheet-like g-C₃N₄, 10 g of urea was placed in a 200 mL closed porcelain crucible. The crucible was then calcinated in a muffle furnace for 4 h at a temperature of 550 °C with a heating rate of 2 °C/min. The black powder of g-C₃N₄ was stored in a glass tube until used in the catalyst fabrication.

2.5. Synthesis of Ce-Doped MIL-88A/g-C₃N₄ Composite

The fabrication process of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ is clarified in Scheme 1. The Ce-doped MIL-88A/g-C₃N₄ composite was prepared as follows: 7.5 mmol of FeCl₃.6H₂O, 2.5 mmol of cerium oxalate, and 10.0 mmol of C₄H₄O₄ were added separately to 50 mL of ultrapure water and stirred for almost an hour at room temperature. Next, a specific mass ratio of g-C₃N₄ ranging from 5 to 20 wt.% was sonicated for 30 min in 10 mL of ultrapure water. The g-C₃N₄ suspension was gradually added over the Ce/Fe/C₄H₄O₄ solution. The resultant mixture solution was agitated for an additional hour, put into a 100 mL autoclave, and heated for 24 h at 85 °C in the oven. The obtained precipitate was separated by centrifuge after naturally cooling to ambient temperature, purified once with ethanol and distilled water, and then left to dry at 85 °C overnight.

2.6. Characterization

The catalysts’ crystal structure and surface functional groups were analyzed using powder X-ray diffraction (XRD-Bruker D8 Advance, Burker, MA, USA), and Fourier-transform infrared spectra (FTIR-Bruker Equinox 55). X-ray photoelectron spectroscopy (XPS-Thermoscientific-Escalab-250Xi VG) investigated the elemental states of samples. The samples’ morphology and microstructure were analyzed using scanning electron microscopy and energy dispersive X-ray microscopy (SEM, SEM-EDX-JSM-760F). The gas chromatography–mass spectrometry (SHIMADZU, GCMS-QP 2010 Ultra) technique investigated the degradation routes.

2.7. Fenton-like Degradation Experiments

TC was degraded in a 50 mL covered container with aluminum foil to avoid the influence of the light on degrading the TC molecules. An amount of 0.01 g of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ was soaked into the TC solution (50 mg/L, 20 mL) with continuous stirring. After attaining the state of adsorption–desorption equilibrium, H₂O₂ was added to the TC-(Ce_0.33Fe) MIL-88A/10%g-C₃N₄ adsorption system. The effects of experimental factors were examined at wide ranges, including the initial concentration of TC (50–300 mg/L), H₂O₂ concentration (10–200 mg/L), temperature (25–55 °C), catalyst dosage (0.005–0.02 g), and pH medium (3–11). During the degradation, a sample was collected and analyzed at intervals times. The TC concentration was investigated via an ultraviolet–visible spectrophotometer at 354 nm. Each experiment was conducted in triplicate to ensure the accuracy and reproducibility of the results. The average value and standard deviation (represented by error bars) were calculated for each set of data. The TC removal rate was calculated as follows:

$$Removal efficiency \%={\displaystyle \frac{{\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_0}}\times 100\%$$

(1)

C₀ = the TC initial concentration, and C_t = the TC concentration at a specific time during the Fenton degradation.

2.8. Quenching Test

Scavengers like tert-butanol (TBA) and chloroform (CF) were added separately to the catalytic system to determine the radical species and define the controlling radicals in the reaction. After adjusting the pH of the solution, 0.01 g of the catalyst, 10 mL of scavenger, and H₂O₂ were added sequentially to 20 mL of TC (50 mg/L). Finally, a sample was taken during the degradation process and examined periodically.

2.9. Regeneration Test

The regeneration test was conducted on the (Ce0_.33Fe) MIL-88A/10%g-C₃N₄ catalyst for five Fenton-like degradation cycles of the TC molecules. After each catalytic cycle, (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ was collected by centrifugation and soaked in 20 mL of NaOH solution with a concentration of 0.01 M for 100 min. Next, the catalyst was separated, washed with distilled H₂O, and dried at 50 °C for reuse in the sequential catalytic cycle.

3. Results and Discussion

3.1. Characterization of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄

3.1.1. Fourier-Transform Infrared Spectra (FT-IR)

FT-IR spectra were applied to explicate the chemical bonding and molecular structure of MIL-88A, (Ce_0.33Fe) MIL-88A, g-C₃N₄, and (Ce_0.33Fe) MIL-88A/10%g-C₃N₄, as illustrated in Figure 1A. The broader absorption band of MIL-88A around 3000–3650 cm⁻¹ was assigned to the O-H stretching vibrations of water adsorbed on the catalyst’s surface. The peaks at 1600.22 and 1396.47 cm⁻¹, respectively, were caused by the asymmetric and symmetric vibrations of the carboxyl group (-COOH) of the fumaric acid [26]. The stretching vibrations of C=O, C-C, Fe-O, and the bending vibration of C-H were responsible for characteristic peaks at 1708.57, 1216.64, 572.23, and 980.29 cm⁻¹, respectively [35]. In the (Ce_0.33Fe) MIL-88A, the redshift of the Fe-O peak to a shorter wavelength (556.20 cm⁻¹) may be explained by the Ce doping and the Ce–O–Fe formation [29]. The relatively large peaks in g-C₃N₄ between 3210.77 and 3062.21 cm⁻¹ are assigned to the bending vibration of terminal NH₂ or NH groups [36]. The belonging peaks of the stretching vibrations of the heterocyclic C-N and C=N groups appeared between the wavenumbers of 1720.02 and 1179.06 cm⁻¹. In addition, the peak at 782.62 cm⁻¹ resulted from the heptazine ring bending vibration [37]. After the combination of (Ce_0.33Fe) MIL-88A with g-C₃N₄, the spectral lines of these composites all match the absorption band of (Ce_0.33Fe) MIL-88A and g-C₃N₄.

3.1.2. X-ray Diffraction (XRD)

The crystallographic characters of MIL-88A, (Ce_0.33Fe) MIL-88A, g-C₃N₄, and (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ were identified by XRD analysis, determining their average crystallite sizes using the Scherrer equation (Figure 1B). For MIL-88A, the diffraction peaks at 2θ = 10.79°, 11.99°, 15.39°, 21.44°, and 12.93°, respectively ascribed by (100), (101), (012), (022), and (110) planes [13,15,38]. Additionally, the average crystallite size of MIL-88A was 5.93 nm, implying its low crystal size and high surface area. On the XRD pattern of (Ce_0.33Fe) MIL-88A, it was noticed that the peaks relating to the (011), (100), and (101) crystal planes appeared at the angles of 10.51°, 11.21°, and 12.70°. In addition, there was a noticeable increase in the crystallite size of (Ce_0.33Fe) MIL-88A (25.43 nm) compared to the pristine MIL-88A, indicating the occurrence of Ce doping. The g-C₃N₄ characteristic peaks were found at 2θ = 19.78° and 29.72°, which corresponded to the diffraction planes of (100) and (002), respectively, with an average crystallite size of 42.55 nm [30]. The diffraction peaks for both (Ce_0.33Fe) MIL-88A and g-C₃N₄ were observed in the crystallographic pattern of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ composite, indicating that the crystal structures were well retained. The average crystallite size (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ was about 64.27 nm, clarifying an increase in the (Ce_0.33Fe) MIL-88A size after decorating with g-C₃N₄ owes to its large crystallite size.

3.1.3. Brunauer–Emmett–Teller (BET)

The nitrogen adsorption/desorption hysteresis loop of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ (Figure 1C) clarified the type (VI) isotherm, disclosing the mesoporous structure of the composite. In addition, the derived specific surface area of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst is 283.21 m²/g. This result reflects the largely available surface area of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst to proceed the Fenton-like degradation process of TC onto its surface.

3.1.4. X-ray Photoelectron Spectroscopy (XPS)

The (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst was analyzed using XPS to identify its chemical components and elemental states, as represented in Figure 2. The wide-spectrum (Figure 2A) revealed that (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ is composed of five elements: C, N, O, Fe, and Ce. The C1s spectrum (Figure 2B) exhibited three peaks at 284.62, 288.67, and 286.33 eV, corresponding to sp² hybridized carbon atoms: C-C/C-H bonding, N=C-N bonding, and C=O groups, respectively [18,39]. The N 1s spectrum (Figure 2C) illustrated the related peaks of the tertiary nitrogen (N–(C)₃) at 400.26 eV and the sp² bonded nitrogen in N-containing aromatic rings (C–N=C) at 399.71 eV [36]. The O 1s peaks (Figure 2D) at 530.65, 531.5, and 531.99 eV are attributed to Fe/Ce-O, C=O/C-O, and water molecule adsorption [40]. The Fe 2p spectrum (Figure 2E) showed the presence of Fe²⁺ 2p3/2 and Fe²⁺ 2p1/2 since the peaks were at 710.79 and 724.04 eV, respectively. Furthermore, the 713.16 and 727.08 eV peaks belong to Fe³⁺ 2p3/2 and Fe³⁺ 2p1/2, respectively. The XPS peaks at 715.65, 719.02, and 732.8 eV are assigned to satellite peaks of 2p3/2 and 2p1/2 [36,41,42]. The Ce 3d peaks (Figure 2F) at 882.38 and 886.02 eV corresponded to Ce³⁺ 3d5/2, while the binding energy at 889.64 and 895.53 eV elucidated Ce⁴⁺ 3d5/2. Moreover, the Ce³⁺ 3d3/2 appeared at 900.67 and 904.3 eV, and the Ce⁴⁺ 3d3/2 manifested at 907.47 eV [43,44].

3.1.5. Scanning Electron Microscopy (SEM)

In the SEM image (Figure 3A), MIL-88A elucidates a hexagonal structure that looks like a spindle with a very smooth surface, demonstrating the development of a fine crystal structure. The (Ce_0.33Fe) MIL-88A catalyst (Figure 3B) displays a similar hexagonal structure to MIL-88A with some differentiation in the roughness of the surface and tiny spread particles on the MIL-88A surface. This change in the surface of MIL-88A indicates doping by cerium particles on the MIL-88A surface. The g-C₃N₄ image shows smooth and aggregate stacking layers that resemble cracked sheets (Figure 3C). Ultimately, the SEM image of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ (Figure 3D) demonstrates that (Ce_0.33Fe) MIL-88A is adhered to the surface of the cracked sheet of g-C₃N₄ and continues to have a hexagonal spindle morphology. The elemental mapping images showed the well-distributed Fe, Ce, C, N, and O elements in the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ composite, as demonstrated in Figure 3E–I. The appearance of the Ce signal indicated the successful doping of cerium species. Additionally, the SEM-EDX pattern denoted that the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ is composed of carbon, nitrogen, oxygen, iron, and cerium with atomic percentages of 23.39%, 2.96%, 64.84%, 7.96%, and 0.85%, sequentially, as depicted in (Figure 3J).

3.2. Fenton-like Degradation of Tetracycline Hydrochloride

3.2.1. Optimization of the Components’ Ratio of Ce-Doped MIL-88A/g-C₃N₄

A series of lab experiments proceeded on TC degradation by Ce-doped MIL-88A/g-C₃N₄ to denote the synergistic effect between Ce species, MIL-88A, and g-C₃N₄ and identify the optimal Ce-doping proportion and the finest g-C₃N₄ amount in the catalyst (Figure 4A–C). The impact of cerium doping on the aptitude of the Fenton-like degradation of TC was investigated by varying the Ce ratio in (Ce_xFe) MIL-88A catalysts in which the TC removal efficiency was 58.82%, 68.89%, 73.01%, 80.40%, and 76.58% when the x ratios were 0.05, 0.1, 0.25, 0.33, and 0.5, respectively. This finding suggested that the finest Ce-doping amount in the Ce-doped MIL-88A is x equaled 0.33. This finding could be explained by the essential role the cerium species in activating H₂O₂ and yielding ROS radicals, where the Ce³⁺ ions could work with the Fe²⁺ions-containing MIL-88A to expand the redox cycle. Furthermore, the optimal proportion of g-C₃N₄ in the (Ce_0.33Fe) MIL-88A/g-C₃N₄ catalysts was determined. The Fenton-like degradation % of TC was 84.49%, 88.78%, 92.44%, and 86.31%, when the g-C₃N₄ proportion was 3%, 5%, 10%, and 20%, respectively. This result clarified that the finest g-C₃N₄ proportion in the (Ce_0.33Fe) MIL-88A/g-C₃N₄ was 10%. This enhancement in the TC degradation efficiency of Ce_0.33Fe) MIL-88A after decorating with g-C₃N₄ owes to the nitrogen’s lone pairs of electrons that can be used as electron donors since g-C₃N₄ itself contains the electron-rich structure of the heptazine ring of the pyridine nitrogen group. So, g-C₃N₄ can easily donate electrons to the to activate H₂O₂ and recover the catalyst’s metal species. Moreover, it was recorded that the Fenton-like degradation efficacy of MIL-88A, (Ce_0.33Fe) MIL-88A, g-C₃N₄, and (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ towards the TC molecules was 52.51%, 80.40%, 21.64%, and 92.44%, respectively. This observation implied the synergetic effect between the Ce species, MIL-88A, and g-C₃N₄ to enhance the produced ROS [18], while in the case of the presence of H₂O₂ alone, the TC elimination rate was only 10.83% after 100 min, indicating that H₂O₂ is inactive as a catalyst for TC degradation.

3.2.2. Effects of pH

The Fenton-like degradation efficiency of TC was studied on a wide scale of the pH of the catalytic medium, as represented in Figure 5A. As the initial pH was elevated from 3 to 7, the degradation % of TC enhanced drastically to 92.44%. Next, raising the pH medium over pH = 7, the degradation efficiency dropped to 60.51% at pH = 11. These results could be attributed to the lower redox potential of H₂O₂ in alkaline media (E^o = 1.8 eV) than in acidic media (E^o = 2.7 eV). Therefore, at a low catalytic pH media, the amount of the produced ^⦁OH is low; furthermore, the high concentrations of H⁺ obstruct the ^⦁OH radicals. Contrariwise, in a low-alkaline and natural media, the higher generated concentrations of ^⦁OH significantly attacked the TC molecules. In addition, the formed hydro per-oxy anions (OOH⁻) in a high-alkaline medium possess higher affinity than H₂O₂ towards the metal species (viz., Ce and Fe), as represented in (Equations (2) and (3)) [45]. Also, auto-decomposition (Equation (4)) is another problem of performing the Fenton-like reaction of TC in a high-alkaline medium [39,46].

H₂O₂ + OH⁻→OOH⁻ + H₂O

(2)

M + OOH⁻→M…^•OOH

(3)

M + OOH⁻→O₂ + 2H₂O

(4)

3.2.3. Effects of Catalyst Dosage

The oxidative degradation of TC by the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst can be boosted by increasing the catalyst dosage, so the Fenton-like degradation of TC was studied at varied doses (Figure 5B). As the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ dosage elevated from 5 to 20 mg/L, the TC elimination rate improved from 74.82% to 99.71% within 100 min. This enhancement can be demonstrated by augmenting the amount of catalyst that promotes the ^⦁OH radicals and increments specific surface area [16].

3.2.4. Effects of Reaction Temperature

The Fenton-like degradation of TC by the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst was examined at various reaction temperatures (Figure 5C). Notably, the TC degradation % diminished from 92.44% to 65.94% as the system temperature elevated from 25 to 55 °C. At low temperatures, TC can be oxidized by ^⦁OH; but, when temperature increased, ^⦁OH became more active and began to self-consume, as elucidated in (Equation (5)). Furthermore, H₂O₂ was unstable at high temperatures and split into water and oxygen [47].

$${\mathrm{H}\mathrm{O}}^{•}+{\mathrm{H}\mathrm{O}}^{•}\to {\mathrm{H}}_2{\mathrm{O}}_2$$

(5)

3.2.5. Effects of H₂O₂ Concentration

The TC degradation efficacy by (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ ameliorated obviously from 74.29% to 92.44% with a rise in the H₂O₂ concentration from 10 to 100 mg/L because of the ample production of ^•OH from the high concentration of H₂O₂ (Figure 5D) [48]. However, the TC removal slightly dropped to 81.45% by elevating the oxidant concentration to 200 mg/L. This catalytic performance may be due to the high concentration of H₂O₂, more than necessary, resulting in scavenged ^•OH and generating the less active HO₂^• radical, as clarified in Equations (6)–(8) [11].

$${\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{H}\mathrm{O}}^{•}\to {\mathrm{H}\mathrm{O}}_2^{•}+{\mathrm{H}}_2\mathrm{O}$$

(6)

$${\mathrm{H}\mathrm{O}}_2^{•}+{\mathrm{H}\mathrm{O}}^{•}\to {\mathrm{H}}_2\mathrm{O}+{\mathrm{O}}_2$$

(7)

$${\mathrm{H}\mathrm{O}}^{•}+{\mathrm{H}\mathrm{O}}^{•}\to {\mathrm{H}}_2{\mathrm{O}}_2$$

(8)

3.2.6. Effects of Initial Concentration

The degradation capability of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ towards various TC concentrations was examined for 200 min, as depicted in (Figure 6A). Obviously, the adsorption process and Fenton-like reaction showed a synergistic effect, where they worked together to efficiently degrade the noxious TC molecules. During the first step (adsorption), the adsorption percent of TC attained 51.96%, 40.45%, 30.76%, and 21.01% after 10 min, when the TC concentrations were 50, 100, 200, and 300 mg/L. Next, by adding H₂O₂ to start the second step (Fenton-like reaction), the TC degradation aptitude reached 97.86%, 84.27%, 77.65%, and 68.15% after 200 min. The results clarified a decline in the degradation efficacy by elevating the TC concentrations, which is most likely due to blocking the active sites of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst, giving rise to the prevention of the generated ROS from being sufficient to remove the excess TC [39].

3.3. Kinetic Study

The TC degradation results in the H₂O₂–(Ce_0.33Fe) MIL-88A/10%g-C₃N₄ system during 200 min was examined using first-order and second-order models (Equations (9) and (10)).

$$\ln {\displaystyle \frac{{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_{\mathrm{o}}}}=-{\mathrm{k}}_1\mathrm{t}$$

(9)

$${\displaystyle \frac{1}{{\mathrm{C}}_{\mathrm{t}}}}={\displaystyle \frac{1}{{\mathrm{C}}_{\mathrm{o}}}}+{\mathrm{k}}_2\mathrm{t}$$

(10)

where t = the reaction time, min, k = the rate constant, min⁻¹, and C_t and C₀ = the final and initial concentrations of TC, mg/L [39].

In light of the derived R² values from the first-order and second-order plots (Figure 6B,C), the Fenton-like degradation of the TC molecules by (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ was best modeled with the first-order model since its R² values of the whole TC concentrations were higher than the R² of the second-order model (

Table 1. Derived parameter of kinetics study of the Fenton-like degradation of TC by (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst.

Kinetic Model	TC Concentrations (mg/L)
	50	100	200	300
First order
k1	0.0137	0.0052	0.0049	0.0036
R2	0.980	0.988	0.988	0.984
Second order
k2	0.1511	0.0157	0.0128	0.0061
R2	0.828	0.973	0.979	0.962

). Furthermore, the k₂ of the TC concentrations of 50, 100, 200, and 300 mg/L were 0.1511, 0.0157, 0.0128, and 0.0061 min⁻¹.

3.4. Identification of Reactive Species

For identifying the reactive radical species within the TC Fenton-like reaction, the quenching experiment proceeded, individually utilizing tert-butanol (TBA) and chloroform (CF) as scavengers for ^•OH and ^•O₂⁻ species (Figure 7A) [49]. The findings of the scavenging experiments showed that adding TBA to the catalytic medium inhibited the degradation percent of TC, denoting the critical function of ^•OH in this process. While adding CF does not affect the TC degradation rate, these results indicate that ^•OH is the leading ROS for TC degradation in the H₂O₂–(Ce_0.33Fe) MIL-88A/10%g-C₃N₄ system.

3.5. Degradation Mechanism

In light of the experimental findings, TC degraded to less noxious molecules throughout two sequential stages: adsorption and the Fenton-like reaction. The XPS survey of the used (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst (Figure 7B) demonstrated an increase in the atomic percentages of nitrogen, carbon, and oxygen, confirmed proceeding the adsorption process of TC (first stage). The adsorption of the TC molecules can occur through the formation of H-bonds between oxygen functional groups between (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ and the hydrogen atoms of TC. However, these H-bonds are weaker than those between water molecules and TC. On the other hand, the coordination bonds consist of the oxygen atoms of TC molecules and the unsaturated metals of the composite (Fe and Ce). Additionally, a π–π interaction was responsible for the adsorption of the aromatic structure of the TC molecule onto the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄. Furthermore, the π–π interaction could contribute to the TC adsorption onto the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ surface since the N and O-containing catalyst can donate lone pairs of electrons to the π-orbitals of the TC molecules [4,50,51].

In the second stage, the degradation of TC occurs through an ^⦁OH radical reaction, according to the results of the quenching experiment, as the presence of radical scavengers decreased the degradation efficiency. The XPS spectra of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ before and after the TC degradation reaction were inspected to demonstrate the possible degradation mechanism. The Fe 2p spectrum of the used (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst exhibits a shift in the Fe 2p3/2 peaks from 710.79 and 713.16 eV to 710.82 and 713.36 eV, respectively. Also, the Fe 2p1/2 peaks demonstrate shifting from 724.04 and 727.08 eV to 724.18 and 716.95 eV, respectively (Figure 7C). Additionally, the Fe³⁺/Fe²⁺ ratio in the used catalyst incremented from 0.484 to 0.597, denoting the participation of Fe²⁺ in the activation of H₂O₂, as clarified in (Equation (11)) [12].

Likewise, the Ce³⁺ species involved in the activation of H₂O₂ (Equation (12)) since the associated peaks to Ce 3d5/2 after the Fenton-like degradation of TC slightly shifted, where the position of the Ce³⁺ peaks shifted from 882.38 and 886.02 eV to 882.6 and 886.12 eV, respectively, and also the belonging peaks to Ce⁴⁺ shifted from 889.64 and 895.53 eV to 890.08 and 895.63 eV, respectively, as clarified in (Figure 7D). Furthermore, the Ce 3d3/2 peaks of the Ce³⁺ species shifted slightly from 900.67 and 904.3 eV to 900.75 and 904.44 eV, and the position of the Ce⁴⁺ peak changed from 907.47 eV to 907.9 eV. Notably, the Ce⁴⁺/Ce³⁺ ratio in the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst elevated from 0.207 to 0.265 after the TC degradation process.

Moreover, g-C₃N₄ possesses the ability to transfer charges, which assists in activating the H₂O₂ since it has the electron-rich heptazine ring structure of the pyridine nitrogen group, so it promotes the production of ^⦁OH radicals. The shifting in the related peaks of nitrogen reflects the contribution of the g-C₃N₄ to the Fenton-like degradation process of TC (Figure 7E). Interestingly, these transferred electrons from g-C₃N₄ could recover the Fe³⁺ and Ce⁴⁺ ions [18,52]. More importantly, the regeneration of Ce³⁺ occurs by Fe²⁺, as elucidated in (Equation (13)), where the redox potential of Fe³⁺/Fe²⁺ (0.77 V) is considerably less than that of Ce⁴⁺/Ce³⁺ (1.44 V). The continuous formation of Fe²⁺ and Ce³⁺ ions by this continuous redox reaction can increase the quantities of ^⦁OH produced by the breakdown of H₂O₂ and enhance the Fenton-like activity [53]. Ultimately, the generated ^⦁OH by activating the H₂O₂ throughout the electron transfer from Fe²⁺, Ce³⁺, and g-C₃N₄ could degrade the TC molecules to less toxic molecules, as clarified in (Equation (14)). Figure 8 summarizes the possible degradation mechanism of the TC molecules via the contribution of adsorption and Fenton-like degradation processes.

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

(11)

$${\mathrm{C}\mathrm{e}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{C}\mathrm{e}}^{4+}+{\mathrm{H}\mathrm{O}}^{•}+{\mathrm{H}\mathrm{O}}^{-}$$

(12)

$${\mathrm{F}\mathrm{e}}^{2+}+{\mathrm{C}\mathrm{e}}^{4+}\to {\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{C}\mathrm{e}}^{3+}$$

(13)

$$\mathrm{T}\mathrm{C}+{\mathrm{H}\mathrm{O}}^{•}\to \mathrm{b}\mathrm{y}\mathrm{p}\mathrm{r}\mathrm{o}\mathrm{d}\mathrm{u}\mathrm{c}\mathrm{t}\mathrm{s}\to {\mathrm{C}\mathrm{O}}_2+{\mathrm{H}}_2\mathrm{O}$$

(14)

3.6. The Degradation Pathway of TC

The GC-MS spectrum (Figure S1) identified the intermediate compounds resulting from the degrading of the TC molecules. The previous studies demonstrated three main mechanisms of degradation: The loss of functional groups, ring-opening and central carbon cleavage, and hydroxylation reactions by adding or substituting ⦁OH. Based on the GC-MS findings, it was suggested the deamidation of TC to form Cpd I, as illustrated in Figure 9. The Cpd I underwent C–N bond cleavage and ring-opening to yield 2,3,8-trihydroxy-4,4a,5,6-tetrahydronaphthalen-1(8aH)-one (Cpd II) and 4,8-dihydroxy-4-methyl-3,4-dihydronaphthalen-1(2H)-one (Cpd III). Further degradation of Cpd II resulted in naphthyl ring-opening component 5-(2-hydroxyallyl) cyclohex-1-enol (Cpd IV) at R.T = 22.94 min, and the partial dihydroxylation compound 4,4a,5,6-tetrahydro-3-hydroxynaphthalen-1(8aH)-one (Cpd V) at R.T = 19.03 min. 5-allylcyclohex-1-enol (Cpd VI) at R.T = 13.61 min was generated by dihydroxylation of the Compound (Cpd IV). Further degradation of Cpd V resulted in the production of 3-(2-methylcyclohex-3-en-1-yl)prop-1-en-2-ol (Cpd VII) and 4-allyl-3-methylcyclohex-1-ene (Cpd VIII) at R.T = 19.81 min through a naphthyl ring opening. The Cpd III underwent naphthyl ring opening, dehydroxylation, and demethylation, producing 2,3-dimethylphenol (Cpd IX) at RT = 7.49 min that furtherly degraded to o-cresol (Cpd X) at RT = 16.08 min [54,55,56].

3.7. Recycling Test

The recycling feature of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst was tested during sequential five Fenton-like degradation cycles of the TC molecules, as elucidated in (Figure 10A). Surprisingly, the degradation % of TC diminished from 85.52% to 72.56% after the 5th catalytic run. This promising result of the recycling test could be implied by the included continuous redox cycle in the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst, endowing it with a self-recovery merit that elongates its lifetime.

3.8. Catalyst Stability

To prove the stability of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ after the regeneration test, the catalyst was analyzed by SEM and SEM-EDX. The SEM Image of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ (Figure 10B) showed no change in the catalyst morphology after the Fenton-like degradation reaction of the TC molecules, implying the promising stability of the as-synthesized Fenton-like catalyst. Furthermore, the SEM-EDX was utilized to define the metal leaching from (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ during the Fenton-like degradation of TC (Figure 10C). The findings demonstrate that the atomic % of the Fe and Ce species in the pure (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ are 7.96 and 0.85%, respectively. After the degradation, a slight metal leaching in Fe and Ce from (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ was recorded, reaching 5.77, and 0.52%, respectively. This observation is most likely due to the six nitrogen lone pairs-containing g-C₃N₄ that can be employed as electron donors to minimize the leaching of metal ions due to their distinct electrical characteristics and N-vacancy capabilities.

3.9. Comparison Study

A comparison study was conducted between the activity (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst and other previously reported Fenton-like catalysts towards degrading the TC molecules, as represented in

Table 2. Comparison study between the catalytic efficiency of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst and previous reported catalysts in relevant studies.

Catalysts	Parameters	Degradation %	Ref.
Fe-MPC	Time = 10 min, pH = 4.3, [H2O2] = 1 mM, [TC] = 40 mg/L, catalyst = 0.02 g.	83%	[57]
Fe3O4-Cs	Time = 120 min, pH = 3, [H2O2] = 10 mM, [TC] = 48.09 mg/L, catalyst = 0.5 g.	96%	[58]
MIL-101(Fe)@CS	Time = 140 min, pH = 5, [H2O2] = 0.09 mol/L, [TC] = 100 mg/L, catalyst = 2 g.	87.36%	[59]
S-mZVI/rGO	Time = 60 min, pH = 5, [H2O2] = 2 mM, [TC] = 20 mg/L, catalyst = 0.2 g.	97.2%	[60]
Fe/Cu-ZSM-5	Time = 10 min, pH = 7, [H2O2] = 1.0 g/L, [TC] = 30 mg/L, catalyst = 0.9 g.	99.02%	[61]
Fe3O4-S	Time = 110 min, pH = 7, [H2O2] = 50 mM, [TC] = 25 mg/L, catalyst = 0.5 g.	82.6%	[62]
CuCo-LDO/CN	Time = 40 min, pH = 6.9, [H2O2] = 15 mM, [TC] = 20 mg/L, catalyst = 0.08 g.	93.2%	[63]
IOT	Time = 360 min, pH = 3.6, [H2O2] = 1700 mg/L, [TC] = 200 mg/L, catalyst = 0.3 g.	95%	[64]
MnFe2O4@HL	Time = 120 min, pH = 3, [H2O2] = 1 mM, [TC] = 0.1 mM, catalyst = 0.3 g.	92.3%	[65]
Cu/CuFe2O4/DE	Time = 90 min, pH = 5, [H2O2] = 4.0 mM, [TC] = 50 mg/L, catalyst = 0.2 g.	95.5%	[66]
(Ce0.33Fe) MIL-88A/10%g-C3N4	Time = 100 min, pH = 7, [H2O2] = 100 mg/L, [TC] = 50 mg/L, catalyst = 0.01 g.	92.44%	This work

. The comparison study clarified the high activity of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst during the Fenton-like degradation of TC using a low catalyst dose compared to the other heterogeneous catalysts.

4. Conclusions

This research developed a novel (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ material as a catalyst for TC degradation via a heterogeneous Fenton-like process. The SEM-EDX, FT-IR, XRD, XPS, BET, and SEM findings showed the morphological, compositional, and structural surface of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄. According to XRD measurements, sharp peaks indicate that the materials generated have crystalline characteristics. The SEM images represented supporting the hexagonal spindle (Ce_0.33Fe) MIL-88A over the surface of the cracked sheet of g-C₃N₄. Under optimal conditions (pH = 7, catalyst dosage = 0.01 g, H₂O₂ = concentration 100 mg/L, temperature = 25 °C, and TC concentration = 50 mg/L), the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄/H₂O₂ system successfully degraded TC by 92.44% within 100 min. In the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄/H₂O₂ system, quenching experiments showed that ^•OH is the principal ROS for TC degradation. The XPS analyses demonstrated a continuous redox cycle inside the active components of (Ce_0.33Fe) MIL-88A/10%g-C₃N₄, producing plenty of electrons for the H₂O₂ activation and the generation of ^•OH. The GC-MS pattern clarified the intermediate compounds from the TC degradation. The SEM-EDX suggested the stability of the (Ce_0.33Fe) MIL-88A/10%g-C₃N₄ catalyst since the cerium and iron leaching concentrations after the Fenton-like degradation are inconsiderable. This finding explained the promising results of the recycling test that clarified a slight dwindle in the degradation % of TC after five catalytic cycles.