Ammonium phosphomolybdate–titanium dioxide composite material as a catalyst for antibiotic degradation under ambient conditions

Abstract

A composite material (designated as AT) of ammonium phosphomolybdate (APM) and titanium dioxide (TiO₂) was synthesized and evaluated for its catalytic ability to degrade tetracycline (TC) via an advanced oxidation process (AOP) under ambient dark conditions in a neutral solution. The catalyst was characterized using BET, XRD, FTIR, XPS, XRF, UV-DRS, and FESEM. Unlike pristine APM, the AT composite eliminated pH constraints in antibiotic degradation. Among various compositions, the 9AT catalyst (APM : TiO₂ = 9 : 1 w/w) exhibited the highest TC removal efficiency. Under optimal conditions (1.2 g L⁻¹ catalyst dose, pH 7.1, and 20 mg L⁻¹ initial TC concentration), 9AT achieved 96.5% TC degradation. Here, TiO₂ acted as a coating/stabilizing material to protect APM, which was only stabilized under acidic conditions and dissolved even in a near-neutral solution. The composite facilitated in situ production of H₂O₂, which generated hydroxyl radicals (˙OH) and singlet oxygen (¹O₂) via Mo⁵⁺ redox cycling, with ¹O₂ being the dominant reactive species. Transformation products (TPs) and plausible degradation pathways were identified. The developed catalyst demonstrated excellent stability, maintaining over 65% efficiency after eight cycles. Additionally, 9AT proved effective in degrading TC in real wastewater, underscoring its practical applicability. Notably, unlike pristine APM, which could be used only under acidic conditions, 9AT remained effective at neutral pH. To the best of our knowledge, this is the first study to fabricate and utilize such a composite as a heterogeneous catalyst for TC degradation under aerobic conditions.

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1. Introduction

In recent times, a wide variety of pharmaceutical contaminants have emerged as environmental pollutants. These substances find extensive applications in human medicine, posing a dual threat to human health and ecological integrity. In this regard, the salient points to underscore are the reduced biodegradability and heightened toxicity of antibiotics in comparison to other organic compounds. The deleterious impact of antibiotics on living organisms is particularly pronounced, even at minimal concentrations.^1,2 Moreover, the imperative to eliminate antibiotics from environmental reservoirs, especially aqueous media, is underscored as a pivotal necessity. Tetracycline (TC), a preeminent antibiotic widely utilized in veterinary medicine, registers discernible concentrations in groundwater, surface water, and wastewater.^3–5 The presence of TC residues in our surroundings has raised significant concerns, given their potential long-term threats to the ecological environment and human well-being.⁶ The stable chemical structure of these residues poses a formidable challenge to current sewage treatment facilities, rendering them unable to achieve complete mineralization.⁷ Compared with conventional remediation techniques such as adsorption, biological treatment, membrane filtration, and ion exchange, photocatalysis and advanced oxidation processes (AOPs) have emerged as highly efficient, cost-effective, and environmentally sustainable strategies for contaminant degradation.⁸ In these advanced methodologies, polyoxometalates (POMs) and their composites stand out owing to their unique multi-electron redox properties and structural versatility, enabling highly effective catalytic degradation of antibiotics such as TC, sulfamethoxazole, and ciprofloxacin.⁹ Notably, these composite materials not only enhance catalytic efficiency but also facilitate the complete mineralization of TC, ultimately transforming it into environmentally benign byproducts.^10,11

TiO₂ is a well-known photocatalyst in the realm of TC degradation.^12,13 However, the inherent limitations of pure TiO₂ are noteworthy: (i) its expansive band gap, measuring 3.2 eV for anatase, confines TiO₂ nanoparticles to UV excitation only; (ii) a pronounced recombination rate of electron–hole pairs, surpassing the kinetics of surface redox reactions, detrimentally affects the quantum efficiency of photocatalytic oxidation; (iii) feeble adsorptive capacities toward organic pollutants lead to an inadequate concentration of the mentioned pollutants in the proximity of the TiO₂ surface. To overcome these challenges, TiO₂-based composites are used for the photocatalytic degradation of environmental contaminants. Research indicates that incorporating metal or nonmetal materials via doping can narrow the band gap, substantially amplifying the catalytic activity across the UV to visible light spectrum without compromising the efficacy.^13,14 Recently, POMs such as ammonium phosphomolybdate (APM) have been introduced as catalysts.¹⁵ All POMs have a multi-electron oxidation state with the central phosphorus atoms. Moreover, they can accept up to six electrons without disrupting the Keggin structure.¹⁶ Notably, the Keggin structure is wholly disrupted in concentrated bases. However, the compounds withstand hydration, dehydration, and heat up to 400–500 °C. Due to its strong impact, high stability, and environmentally favorable features, different materials such as carbon,^17–19 β-cyclodextrin,²⁰ polyaniline,²¹ iron,^23,24 zinc,^25,26 gold,²⁷ and TiO₂ (ref. 28 and 29) have been widely incorporated with a polyoxometalate to produce effective polyoxometalate composites for the removal and mineralization of TC antibiotics.

Although TiO₂ is widely known as a photocatalyst, in this study it serves as a coating matrix for APM in the AT catalyst, enabling redox-mediated degradation of TC under dark aerobic conditions at near-neutral pH of solutions. This study focused on synthesizing a novel TiO₂-decorated APM catalyst toward efficient TC degradation for the first time. It is worth mentioning that no research has yet been reported in the literature for the decomposition of TC under ambient dark conditions using TiO₂ anchored on APM (marked as AT). To the best of our knowledge, little information is available on the mechanism, characterization, and environmental application of the AT composite, as well as its capability to contribute to the degradation of TC. Various operational parameters including temperature, concentrations of TC, dosages of AT, and pH levels were examined to see their impacts on the oxidative degradation of TC. By utilizing MALDI-ToF-MS, the study successfully identified and characterized the degradation intermediates of TC within the AT composite system, thereby proposing potential pathways for TC breakdown. Furthermore, the study findings strongly suggest that the AT-oxidation-catalytic system shows substantial promise in efficiently eliminating TC pollutants from authentic wastewater matrices, revealing its potential practical application.

2. Experimental section

2.1. Materials

TC was purchased from HiMedia Laboratories Pvt. Ltd. Ammonium heptamolybdate [(NH₄)₆Mo₇O₂₄·4H₂O], sodium dihydrogen phosphate (NaH₂PO₄·2H₂O), sodium hydroxide (NaOH), sodium chloride (NaCl), sodium carbonate (Na₂CO₃), sodium bicarbonate (NaHCO₃), sodium nitrate (NaNO₃), sodium sulfate (Na₂SO₄), magnesium sulfate (MgSO₄), manganese sulfate (MnSO₄·H₂O), iron sulfate (FeSO₄·7H₂O), copper sulfate (CuSO₄·5H₂O), potassium hydrogen phthalate (C₈H₅KO₄), mercury sulfate (HgSO₄), silver sulfate (Ag₂SO₄), sulfuric acid (H₂SO₄), titanium dioxide (TiO₂), tert-butyl alcohol ((CH₃)₃COH), hydrochloric acid (HCl), nitric acid (HNO₃), potassium iodide (KI), ethyl acetate (CH₃COOC₂H₅), benzoquinone (BQ), and L-histidine used in the experiment were obtained from Merck, India. Sigma-Aldrich provided 2,2,6,6-tetramethylpiperidine (TEMP), catalase, and 5,5-dimethyl-1-pyrroline N-oxide (DMPO). Unless specified, all other reagent-grade chemicals and solvents were obtained from Merck, India. All solutions in the study were prepared with deionized water obtained using Millipore Elix (Germany). The pH of each solution was adjusted by using 0.1 N NaOH or 0.1 N HCl solutions.

2.2. Catalyst synthesis

In this work, we synthesized a new AT composite to eliminate the pH barrier for antibiotic degradation by microporous APM. The AT composite was prepared by mixing the dissolved salt precursor solution of APM (ammonium heptamolybdate and sodium dihydrogen phosphate) and suspended commercial TiO₂ powder in distilled water (DI), according to the dip-coating process.³⁰ Briefly, 0.1 g TiO₂ powder was suspended in 200 ml DI and mixed at 80 °C for 60 min using a magnetic stirrer at 500 rpm. Since the mass proportion of APM and TiO₂ in the 6AT composite was 6 : 1, 0.6 g of APM salt precursor was added (for 6AT) to the TiO₂ suspension. This solution contained ammonium heptamolybdate and sodium dihydrogen phosphate, which were dissolved in 50 mL of DI in a stoichiometric molar ratio at 80 °C. After addition, the mixture was stirred continuously for an additional 30 min at the same temperature and stirring rate. Subsequently, 6 mL of concentrated HNO₃ (∼70%) was added dropwise to acidify the solution, and the mixture was stirred for another 30 min. The solution was then maintained at room temperature for an additional 30 min to facilitate precipitation. The white 6AT composite was precipitated and washed multiple times with DI to get its pH approximately equal to the pH of DI. Then, the precipitate was dried in an oven at 60 °C to furnish the product. In this way, three more types of AT composites, namely, 8AT, 9AT, and 10AT, with varying mass proportions, viz., APM : TiO₂ (w/w) = 8 : 1, 9 : 1 and 10 : 1, respectively, were also prepared. For the preparation of the 10AT composite, a whitish-yellow precipitate was obtained, revealing that TiO₂ particles did not thoroughly coat APM particles. To ensure the reproducibility and robustness of the synthesis method, each AT composite was prepared multiple times (more than five independent syntheses), and in every case, the resulting catalyst exhibited highly consistent degradation efficiency under identical experimental conditions. This highlights the stability and repeatability of the composite preparation protocol, making it suitable for practical and scalable applications.

2.3. Instrumentation

The examination of the elemental composition and surface properties of the catalyst requires the use of sophisticated analytical instruments. X-ray fluorescence (XRF) (PANalytical, Almelo, Netherlands) was used to determine the elemental composition of the catalysts. To learn more about the chemical makeup of samples, X-ray photoelectron spectroscopy (XPS) analysis was performed. An electron spectrometer (ULVAC PHI 5000 Versa Probe III, USA) was used to capture the XPS spectra of the materials. A high-resolution microscope (Merlin, Carl Zeiss-Gemini FESEM 2, Germany) equipped with an Oxford EDS was used to examine the surface morphologies and elemental composition of catalysts. The crystal structures of the samples were analyzed by powder X-ray diffraction (XRD) using a high-resolution instrument (X'Pert Pro MPD PW3050/60, PANalytical, Almelo, Netherlands). A Cu Kα₁ radiation (wavelength: λ = 1.5406 Å) generator working at 45 kV and 40 mA with a range of scattering angles (2θ) from 10° to 70° and a step size of 0.02° was used. The XRD data were analyzed using the ICDD card. The textural properties (specific surface area, mean pore diameter, and pore volume) of the materials were analyzed using an Autosorb iQ (Quantachrome Instruments, USA) at −196 °C (77 K) to obtain the results of Barrett–Joyner–Halenda (BJH), density functional theory (DFT), and Brunauer–Emmett–Teller (BET) plots. Before each adsorption–desorption evaluation, the samples were degassed at 150 °C (423 K). Fourier transform infrared (FTIR) spectroscopy was performed using an instrument (Nicolet 6700, Thermo Fisher Scientific Instruments, USA) in the scanning range of 400–4000 cm⁻¹. The DRS (diffuse reflectance spectra) of the materials were recorded using a UV-visible spectrophotometer (Model-Cary 5000, Agilent Technologies).

2.4. Experimental procedure

2.4.1. Degradation of TC in the presence of AT as a catalyst under aerobic conditions. A specific volume of TC solution and catalyst at a certain dose was introduced into a series of 15 mL vials placed on an orbital shaker (ACMS Technocracy (P) Ltd., Delhi) under precise temperature control, swirling gently at 150 rpm. An aliquot of 3 mL samples was drawn from their designated vials at scheduled time intervals using high-precision 0.22 μm MCE syringe filters. Subsequently, the absorbance values of the solutions were measured at 356 nm using a UV-visible spectrophotometer (Model-Cary 60, Agilent Technologies). Using a calibration equation, the remaining TC concentrations were determined. Assessments for degradation were carried out in authentic water samples containing TC at 20 mg L⁻¹ concentration, with an initial pH of 7.1, a reaction duration of 120 min, a temperature set at 30 °C, and a catalyst dose of 1.2 g L⁻¹. Each procedure was performed three times, and the results are presented as average values in the appropriate graphical depictions. All degradation experiments were conducted in an orbital shaker under completely dark ambient conditions. No UV or visible light source was applied.

2.4.2. Determination of TC in real water matrices. Pond water and a secondary effluent from the wastewater treatment plant (WWTP) at IIT Kharagpur campus were used as natural surface water and real wastewater, respectively. The process for degrading TC in these waters follows the methodology outlined in section 2.4.1, with the distinction that the TC solution was formulated using real water matrices. The ion chromatography (IC, Metrohm, Switzerland) technique measured the cation and anion concentrations within the real wastewater matrices. The dissolved oxygen level was also gauged using a DO Meter (2FD354, Germany). For the identification of resulting transformation products and the proposal of a degradation mechanism for TC, the MALDI-ToF/ToF technology (Ultraflextreme, Bruker Daltonik GmbH Life Sciences, Germany) was employed to determine the molecular masses of the by-products involved in the process.

2.4.3. Determination of TC. The TC concentration was calculated using a calibration curve equation: absorbance = 0.0318 × conc. (mg L⁻¹) + 0.0425 (R² = 1) at λ_max = 356 nm (limit of detection = 0.062 mg L⁻¹ and limit of quantification = 0.19 mg L⁻¹). The degradation kinetics of TC was simulated by the first-order kinetic model in eqn (1). The closed reflux colorimetric method was used to calculate the chemical oxygen demand (COD).³¹ Analysis of the total organic carbon (TOC) content in the sample was carried out using an Aurora O-I-Analytical TOC analyzer. The degradation efficiency for TC, COD, and TOC removal efficiency are expressed using eqn (2):

ln(C_t/C₀) = −k·t (1)

R = (1 − C_t/C₀) × 100% (2)

where k is the rate constant of the first-order kinetic model (min⁻¹), R is the efficiency of TC degradation or COD or TOC removal (%), C₀ is the initial concentration of TC (mg L⁻¹) or COD or TOC of feed (mg L⁻¹), C_t is the concentration of TC or COD or TOC in the treated sample at time t, and t is the reaction time (min).

2.4.4. EPR measurement. Electron Paramagnetic Resonance (EPR) spectroscopy was performed to examine singlet oxygen (¹O₂) and hydroxyl radical (˙OH) generated by the 9AT catalyst under dark ambient conditions. To detect ¹O₂, a homogeneous mixture of 1.2 mg of 9AT and 20 μL of TEMP was diligently blended in 1 mL of ethyl acetate and stirred for 5 min. This concoction was promptly extracted using a 100 μL capillary tube and inserted into the EPR cavity of a Bruker ELEXSYS 580 spectrometer. X-band spectra were then recorded at room temperature, employing 9.64 GHz microwave frequency, 15 mW microwave power, and 100 kHz modulation frequency. For the detection of ˙OH radicals using EPR, a similar procedure was followed, except that the mixture of 1.2 mg of 9AT and 20 μL of DMPO was uniformly mixed in 1 mL of DI, similarly sampled and subjected to EPR analysis.

3. Results and discussion

3.1. Characterization of catalyst

3.1.1. Structural and chemical characterization. The XRD patterns of 6AT, 8AT, 9AT, and 10AT composites and their parent compounds (APM and TiO₂) are depicted in Fig. 1. The XRD patterns of the parent compounds are indexed with ICDD files for the APM cubic phase and TiO₂ anatase phase. The peaks appearing at 2θ around 25.24°, 37.73°, 47.98°, 53.83°, 55.00°, 62.63°, and 68.70° elucidate the diffractions of the (101), (004), (200), (105), (211), (204), and (116) planes (ICDD card no. 00-083-2243), respectively, of anatase-type TiO₂.^32,33 As the mass proportion of APM increases in the AT composites, a gradual transition in the XRD pattern is observed. The diffraction peaks shift from those characteristic of the anatase phase of TiO₂ to those corresponding to the cubic crystalline phase of APM (ICDD card no. 00-009-0412). This structural transformation indicates the progressive incorporation of APM into the composite matrix, influencing its overall crystallographic properties. Moreover, the determination of vital structural characteristics for the as-prepared APM sample involves the analysis of lattice parameter (α), crystallite size (D), dislocation density (δ), microstrain (ε), and crystallinity index (CI), as delineated in Table 1. The extraction of these parameters was facilitated by utilizing the XRD data, which were meticulously processed using the Origin software. The CI computation was accomplished using the XRD peak height method, assuming a Gaussian function for each peak. For a comprehensive understanding of the mathematical formulations of each parameter, refer to Text S1 (section 1.1) in the ESI† of our previous paper.¹⁵ The CI is increased with the increase in the relative amount of APM (Table 1).

Fig. 1 XRD spectra of APM, TiO₂, and corresponding composites.

Table 1 Average estimated values of the lattice parameter (α), crystallite size (D), dislocation density (δ), microstrain (ε), and crystallinity index (CI) corresponding to the XRD patterns of different APM–TiO₂ materials

Materials	Lattice parameter α (Å)	Crystallite size D (nm)	Dislocation density δ × 10^10 (line per cm^2)	Microstrain ε × 10^−3	Crystallinity index CI (%)
APM	11.70	31.55	10.71	4.49	79.65
TiO2	6.59	20.17	20.74	4.18	20.17
6AT	7.36	24.73	16.85	4.25	63.34
8AT	9.88	25.24	16.43	4.31	66.53
9AT	10.08	25.65	16.10	4.36	70.63
10AT	11.04	27.64	13.78	4.44	73.54

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The FTIR analysis unveiled the surface functionalities of the samples. Fig. 2 illustrates the distinctive features, particularly in the low wavenumber range of 1100–500 cm⁻¹, indicative of the Keggin-type structure formation in all composite materials.¹⁵ It is noteworthy that the spectral elements include a broad peak at 3207 cm⁻¹ and a sharp band at 1411 cm⁻¹, attributed to the absorption of the –NH₄⁺ unit.³⁴ A characteristic band at 595 cm⁻¹ corresponds to the P–O bending vibration (ν_b) frequency.³⁴ The peaks observed at 3651, 3425, and 1625 cm⁻¹ are associated with the stretching (ν_s) and bending (ν_b) vibration frequencies of the O–H group, respectively.^15,34 In contrast, the FTIR spectrum of the commercial TiO₂ sample (Fig. 2) displays the stretching (ν_s) bands of the O–H group in the range of 3730 to 3500 cm⁻¹.³⁵ All the composite materials mostly exhibit the characteristic peaks of APM and weakly of TiO₂, affirming the successful synthesis of the heterostructure.

Fig. 2 FTIR spectra of APM, TiO₂, and corresponding composites.

The catalysts underwent a comprehensive chemical characterization via XPS analysis, with all spectra calibrated against the adventitious carbon peak at 284.8 eV. The XPS survey spectra, showcased in Fig. S1,† illustrate the presence of various elements in both pristine and utilized 9AT composites. The high-resolution Mo 3d spectra (Fig. 3) of the pristine 9AT reveals distinctive peaks at 232.8 and 235.9 eV, corresponding to 3d_5/2 and 3d_3/2, respectively, aligning with the molybdenum(VI) state reported in the literature.^15,36 In contrast, post-TC degradation, the Mo 3d region exhibited peaks at 231.3, 233 eV, and 234.4, 236.1 eV, assigned to 3d_5/2 and 3d_3/2, respectively. Notably, the 233 and 236.1 eV peaks indicate the Mo(VI) state, while the 231.3 and 234.4 eV peaks signify the molybdenum(V) form.^15,36 The Ti 2p peaks at 459.3 and 464.9 eV (Fig. S2a†) confirm Ti in its (IV) form in both pristine and used 9AT composites.^32,35 The deconvolution of the N 1s region reveals a peak at 398.4 eV, corresponding to –NH– for both states of 9AT (Fig. S2b†).¹⁵ The O 1s high-resolution spectra exhibit a distinct peak at 530.6 eV (Fig. S2c†), attributed to the characteristic peaks of metal oxide (MoO₃).^15,36 High-resolution P 2p spectra display prominent peaks at 133.5 and 134.4 eV (Fig. S2d†), corresponding to 2p_3/2 and 2p_1/2.^15,37 The XPS findings underscore variable Mo oxidation states (between V and VI) and consistent oxidation levels in other elements during TC degradation. Despite peak shifts, the 3.1 eV peak-to-peak energy difference (Fig. 3) indicates the stability of Keggin units throughout the TC breakdown, emphasizing the pivotal role of 9AT in catalyzing TC oxidation via a cyclic redox mechanism involving Mo(VI) and Mo(V) state interconversion.

Fig. 3 High-resolution XPS spectra of the pristine and reacted 9AT composite after TC degradation for Mo 3d.

3.1.2. Morphological analysis. Surface morphology assessments are conducted for the APM, commercial TiO₂, 6AT, 8AT, 9AT, and 10AT samples. The nitrogen adsorption–desorption isotherms illustrated in Fig. 4(a–f) reveal intriguing patterns. In Fig. 4b, TiO₂ demonstrates a characteristic type IV BET isotherm, indicative of a mesoporous texture with a pore radius measuring 30.44 Å (Table 2). In contrast, APM and all AT composites exhibit a distinctive type I BET isotherm, signifying the presence of micropores.³⁸ The microporous nature is corroborated by the average pore radius data detailed in Table 2, derived from the BET, DFT, and BJH plots. Micropore volume, micropore area, and external surface area for all samples are succinctly presented in Table 3, determined through the V–t plot method. Notably, an increase in APM within the composite materials corresponds to the heightened surface area and pore volume, suggesting an enhanced adsorption capacity for the catalyst.

Fig. 4 BET isotherm of the (a) APM, (b) commercial TiO₂, (c) 6AT, (d) 8AT, (e) 9AT, and (f) 10AT samples.

Table 2 Textural properties of different APM–TiO₂ materials

Materials	Methods	Specific surface area (m^2 g^−1)	Total pore volume (cm^3 g^−1)	Average pore radius (Å)
APM	BET	193.88	0.101	10.39
	DFT	267.86	0.096	4.51
	BJH	12.73	0.018	18.12
TiO2	BET	11.54	0.018	30.44
	DFT	10.89	0.016	17.68
	BJH	8.35	0.013	15.45
6AT	BET	62.71	0.027	19.83
	DFT	98.22	0.031	8.64
	BJH	8.87	0.014	16.05
8AT	BET	84.53	0.044	11.22
	DFT	112.48	0.048	5.21
	BJH	9.16	0.015	17.05
9AT	BET	107.78	0.061	10.41
	DFT	160.16	0.059	4.55
	BJH	10.42	0.016	18.07
10AT	BET	144.76	0.077	10.40
	DFT	213.62	0.063	4.52
	BJH	11.24	0.017	18.10

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Table 3 Micropore volume, micropore area, and external surface area of different APM–TiO₂ materials as determined using the V–t plot method

Materials	Micropore volume (cm^3 g^−1)	Micropore area (m^2 g^−1)	External surface area (m^2 g^−1)
APM	0.062	147.67	46.21
TiO2	0.00	0.00	11.54
6AT	0.011	7.23	21.36
8AT	0.024	29.35	28.12
9AT	0.033	72.52	35.26
10AT	0.051	102.48	41.23

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SEM-EDS is a valuable tool for discerning the surface morphology and chemical compositions of materials. The visual analysis of the APM within Fig. 5a reveals a porous structure with a distinctive rhombododecahedral morphology, as detailed in prior research.¹⁵ Conversely, the SEM depiction in Fig. 5b showcases uniformly distributed spheres, attesting to the uniformity and compactness of the commercial TiO₂ particles. Moreover, the crystallite size of TiO₂, as corroborated by the data in Table 1, is approximately 20 nm. Noteworthy is the evident size discrepancy between APM particles and pure TiO₂ particles, as depicted in Fig. 5a and b. From Fig. 5c, it is evident that TiO₂ agglomerates across the entire surface of APM, facilitated by the interaction with microporous APM particles. The subsequent Fig. 5c and d illustrate a noticeable reduction in TiO₂ coverage on APM particles as the APM content increases in the composite materials. Specifically, the 10AT composite, shown in Fig. 5d, exhibits a decreased extent of TiO₂ coating on APM particles, suggesting variations in the material's surface characteristics and interaction dynamics. The oxide ratio of APM/TiO₂ was elucidated through EDS analysis (refer to Table 4), while XRF analysis (also presented in Table 4) identified the primary oxides of Ti, Mo, N, and P, confirming the successful coating of TiO₂ particles on APM particles.

Fig. 5 FESEM image of the (a) APM, (b) commercial TiO₂, (c) 6AT, (d) 8AT, (e) 9AT, and (f) 10 AT samples.

Table 4 Oxide compositions of different APM–TiO₂ materials

Oxides	APM		TiO2		6AT		8AT		9AT		10AT
	EDS	XRF	EDS	XRF	EDS	XRF	EDS	XRF	EDS	XRF	EDS	XRF
MoO3	52.78	53.46	0.00	0.00	44.12	45.78	47.21	47.43	47.71	47.88	48.71	48.67
N2O5	39.71	39.54	0.00	0.00	32.67	33.88	35.11	35.09	34.42	35.41	35.35	35.41
P2O5	7.51	7.00	0.00	0.00	5.98	6.00	6.24	6.21	6.21	6.29	6.37	6.35
TiO2	0.00	0.00	100.0	100.0	17.23	14.34	11.44	11.27	11.66	10.42	9.57	9.07

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3.1.3. Zero-point charge and optical characterization. The isoelectric point, also known as the point of zero charge (pH_ZPC), plays a pivotal role in heterogeneous catalysis. The pH_ZPC of the catalysts APM, commercial TiO₂, 6AT, 8AT, 9AT, and 10AT composites were determined to be 3.07, 6.67, 6.65, 6.61, 6.58, and 4.01, respectively. Notably, the lower pH_ZPC of the 10AT composite again implies an incomplete coating of APM particles by TiO₂. The UV-DRS of APM, TiO₂, and APM/TiO₂ composites with varying APM loading (depicted in Fig. 6a) showcase a broader absorption range for APM and all AT composites compared to the narrow response of commercial TiO₂ below 400 nm. This analysis underscores the effectiveness of APM and AT composites across all studied regions. The UV-visible DRS analysis was performed to understand the electronic structure of the AT composite. The electronic structure may influence the redox activity of the AT composite even under dark conditions. Increasing the APM content induces the redshifts of absorption bands in the UV-visible spectrum of these composite materials. Furthermore, the theoretical expression for the determination of the band gap of a material is given as follows:

αhν = K(hν – E_g)^1/2

where α is the absorption coefficient, K is a constant, hν is the discrete photo energy, and E_g is the band gap energy. A classical Tauc approach was further employed to estimate the E_g value of the material.³⁹ The plot of “(αhν)² in (a.u.)” versus “photon energy (hν) in eV” based on the direct transition and extrapolation of hν at α = 0 gives the absorption edge energy, which is the band gap of the materials (Fig. 6b). The band gap energies of APM, TiO₂, 6AT, 8AT, 9AT, and 10AT composites are 2.43, 3.20, 2.55, 2.51, 2.48, and 2.46 eV, respectively. Remarkably, the decrease in band gap energy with the increase in APM content suggests generating new energy levels.

Fig. 6 (a) UV-DRS spectra of different catalysts and (b) the corresponding direct transition Tauc plot for band gap estimation.

3.2. Catalytic performance

The investigational exploration into the catalytic efficacy of the freshly prepared catalysts centered on the degradation of TC antibiotics under dark aerobic conditions (Fig. 7a). To investigate the reproducibility, triplicate experimental runs were conducted under identical conditions for each data point, and the mean values (with the error bars) were plotted. The standard deviations in all cases were found to be consistently below ±0.5, reflecting the high repeatability and precision of the measurements. The specific first-order rate constant (k) and coefficient of determination (R²) are compiled in Table 5. The visual representation in Fig. 7a highlights that the TiO₂ sample demonstrated a meager removal efficiency of approximately 5% after a 120 min reaction period. It is well established that APM becomes soluble at near-neutral pH (∼7).¹⁵ However, the degradation efficiency of pure APM under similar conditions has been examined for comparison. Under dark ambient conditions, the microporous APM material exhibited a noteworthy ability to degrade 31.2% of TC. Notably, the AT composites outperformed their parent materials in terms of degradation efficiency. The 6AT, 8AT, 9AT, and 10AT samples showcased degradation efficacies of 78.1%, 88.5%, 96.5%, and 41.1%, respectively, following a 120 min reaction. The enhancement in degradation performance initially correlated with the increase in APM content, reaching its optimum at a mass ratio of 9/1 in the APM/TiO₂ composite. This improved performance can be attributed to the combined effects of increased surface area and enhanced catalytic activity. However, a further increase in the APM content beyond this optimal ratio led to a decline in degradation efficiency. The 10AT composite exhibited lower performance, probably due to alterations in material composition and structural characteristics. Based on these findings, 9AT was selected as the optimal catalyst for further studies on TC degradation. This optimality arises from a synergistic balance among high surface area, microporosity, and catalytic reactivity conferred by the Mo(VI)/Mo(V) redox cycle of APM, while the TiO₂ component enhances the structural stability and prevents the dissolution of APM at near-neutral pH. A further increase in APM content (as in 10AT) disrupts this balance, leading to a decline in performance due to incomplete TiO₂ coating and compromised composite integrity.

Fig. 7 (a) Performance of the APM, TiO₂, 6AT, 8AT, 9AT, and 10AT catalysts under dark ambient conditions; alteration of the degradation efficiency of 9AT at different (b) initial pH values, (c) 9AT dosages, (d) TC concentrations, and (e) temperatures; and (f) first-order reaction kinetics for different temperatures (reaction conditions: initial TC concentration = 20 mg L⁻¹, catalyst = 1.20 g L⁻¹, temperature = 30 °C, agitation = 150 rpm under neutral initial pH (7.1), except for the tested condition).

Table 5 First-order rate constant (k) values for TC degradation under different conditions

Conditions	Rate constant (k) (min^−1)
TC = 20 mg L^−1Catalyst = 1.2 g L^−1 Initial pH = 7.1 Temperature = 30 °C Time = 120 min Agitation = 150 rpm (except for tested conditions)	Catalyst
	APM		TiO2		6AT
	k	R ^2	k	R ^2	k	R ^2
	0.003	0.99	0.0004	0.95	0.011	0.95
	Catalyst
	8AT		9AT		10AT
	k	R ^2	k	R ^2	k	R ^2
	0.017	0.98	0.024	0.94	0.004	0.95
	Initial pH
	3.1		7.1		8.3
	k	R ^2	k	R ^2	k	R ^2
	0.018	0.95	0.024	0.94	0.020	0.95
	9AT dose (g L^−1)
	1.0		1.2		1.4
	k	R ^2	k	R ^2	k	R ^2
	0.009	0.98	0.024	0.94	0.024	0.93
	TC concentration (mg L^−1)
	5		10		20		30
	k	R ^2	k	R ^2	k	R ^2	k	R ^2
	0.04	0.98	0.027	0.96	0.024	0.94	0.006	0.96
	Temperature (°C)
	25		30		35
	k	R ^2	k	R ^2	k	R ^2
	0.017	0.95	0.024	0.94	0.031	0.94
	Real wastewater matrices and real TC tablet
	Pond water		Real wastewater		Real TC tablet in DI
	k	R ^2	k	R ^2	k	R ^2
	0.017	0.95	0.016	0.97	0.027	0.96

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The significance of solution pH in the degradation process is pivotal, as the stability of APM in an acidic medium contrasts with its dissociation at pH levels exceeding 7. Notably, complete dissociation occurs beyond pH 11, forming soluble molecular precursors, including NH₄⁺, PO₄⁻³, and [Mo_xO_y]^−z ions.¹⁵ Investigation into the impact of solution pH on TC degradation performance by 9AT involved adjustments within the pH range of 3.1–8.3, with the results illustrated in Fig. 7b. Given the unique functional groups of the TC molecule, characterized by distinct pK_a values of 3.3, 7.7, and 9.7 (shown in Fig. S3†), its existence manifests as cationic (TC⁺) at pH < 3.3, zwitterionic (neutral TC^±) within the pH range of 3.3–7.7, and anionic (TC⁻ and TC⁻²) at pH > 7.7.¹⁵Fig. 7b portrays the removal efficiencies exceeding 91% across the initial pH range of 3.1–8.3. Notably, the optimal degradation efficiency, reaching 96.5%, was achieved at a solution pH of 7.1, corresponding to the zwitterionic form of TC. Thus, pH 7.1 emerges as the selected optimum pH for subsequent experiments in this study. This is in contrast to our earlier work on TC degradation carried out in the acidic range of pH (∼4.1) using APM alone.¹⁵

The optimization of catalyst (9AT) dosage is integral to enhancing practical applications, as evidenced by a series of experiments conducted within the dosage range of 1.0–1.4 g L⁻¹. The outcomes, shown in Fig. 7c, vividly portray a significant surge in catalyst efficacy, escalating from 67.1% to 96.5% after a 120 min reaction period when the dosage is elevated from 1.0 g L⁻¹ to 1.2 g L⁻¹. However, intriguingly, further increments in dosage failed to yield additional improvements in catalyst performance. The initial dosage escalation proved instrumental in augmenting the abundance of active reaction sites, thereby expediting the degradation process. Consequently, the dosage of 1.2 g L⁻¹ emerged as the optimal point, a judicious choice that has been embraced for subsequent studies, ensuring an economically viable procedure.

Moreover, the research delved into the impact of different initial concentrations of TC, spanning from 5 to 30 mg L⁻¹, on the degradation of TC. As depicted in Fig. 7d, the efficiency of TC removal reached its pinnacle at 98% within a 90 min timeframe when the initial concentration was set at 5 mg L⁻¹. In contrast, this effectiveness markedly dwindles to 56.2% within 120 min when the initial concentration is heightened to 30 mg L⁻¹. This decline can be ascribed to the diminished accessibility of active sites on the APM surface and the generation of fewer reactive species, consequently impeding the degradation process at elevated initial TC concentrations.

The influence of reaction temperature extends its sway over both the pace of chemical reactions and the generation rate of radicals. As illustrated in Fig. 7e, the effectiveness in breaking down TC heightened with rising temperatures. Notably, a significant upsurge in the removal rate of TC was noted as the temperature escalated from 25 to 35 °C. In 120 min, the TC elimination rate reached its peak at 98.8% at 35 °C, strongly indicating the inclination of TC degradation in correlation with temperature elevation. This surge aligns with the principle that as the temperature increases (expressed by the equation k = A exp^(−E_a/RT)),⁴⁰ the pace of chemical reactions accelerates. Elevated temperatures promote more efficient molecular collisions and activation, thereby enhancing the average energy of activated molecules while reducing the empirical activation energy, E_a. Consequently, this environment becomes more conducive to the unfolding of reactions. The disintegration kinetics of TC closely adheres to the first-order model, as portrayed in Fig. 7f. Furthermore, the augmentation in temperature expedites the generation of active compounds, thereby nurturing the degradation process of TC. The 120 min reaction time was selected as the optimal operational window based on the kinetic analysis and first-order degradation behavior, where near-complete TC degradation and transformation of intermediates are consistently achieved. While shorter durations may suffice under specific conditions, the 120 min benchmark ensures robust performance across varying concentrations and real water matrices.

3.3. Catalyst reusability

An extensive eight-cycle recycling investigation was conducted during the degradation of TC to assess the reusability of the as-synthesized 9AT as the catalyst, as depicted in Fig. 8a. Each catalytic reduction cycle ends with the addition of a new batch of tetracycline solution to the reaction mixture. As illustrated in Fig. 8a, the 9AT composite exhibits commendable catalytic efficacy, persisting through eight cycles. The TC removal percentages are 90.7, 88.8, 85.7, 83.1, 76.6, 74.1, 69.2, and 65.7 for the eight consecutive recycles, underscoring the high stability and reusability of the 9AT composites. This decline is attributed to the reduced molybdenum (Mo⁺⁵) production and a layer of small particles from fragmented TC molecules, which impede the oxidative activity of the 9AT catalyst.^15,41,42 The original activity of the 9AT composite is restored with HNO₃ treatment as the treated 9AT composite has removed 94.3% TC. Concentrated HNO₃, functioning as a potent oxidant, transforms Mo(V) to Mo(VI), thereby restoring the catalytic activity of the 9AT composite. Furthermore, the XRD spectrum of the recycled catalyst was scrutinized to ascertain the stability of the composite. In Fig. 8b, the XRD spectra reveal that the catalyst maintained its structure after the eight-cycle degradation experiment. A comparative analysis of the FTIR spectra (Fig. 8c) between the pristine and reused material substantiated that the catalyst retained its structural integrity after the recycling experiment. The FESEM images (Fig. 8d) further underscored the exceptional stability of the reused material. In summation, the catalyst exhibits commendable reusability in TC degradation.

Fig. 8 (a) Cyclic TC degradation performance of the 9AT composite, (b) comparison between the XRD spectrum of pristine and 8th-time recycled 9AT sample, (c) comparison between the FTIR spectrum of the pristine and 8th-time recycled 9AT samples, and (d) FESEM image of the 8th-time recycled 9AT composite (reaction conditions: initial TC concentration = 20 mg L⁻¹, 9AT = 1.2 g L⁻¹, temperature = 30 °C, agitation = 150 rpm under neutral initial pH (7.1)).

3.4. Identification of reactive oxygen species in the reaction medium

To unravel the specific reactive oxygen species (ROS) orchestrating the degradation of TC employing the 9AT composite as a catalyst, a set of six widely recognized quenchers, namely, benzoquinone (BQ), catalase, ethanol (EtOH), L-histidine, potassium iodide (KI), and tert-butyl alcohol (TBA) are enlisted. Each quencher, except BQ, was introduced at a concentration of 10 mM, while BQ featured at 1 mM. The established functions of these quenchers, attributing BQ, EtOH, L-histidine, KI, and TBA as quenchers for ˙O₂⁻, ˙OH, ¹O₂, ˙OH_surface, ˙OH_free, respectively, have been documented in prior studies.¹⁵ Catalase was used to quench the in situ-generated H₂O₂.⁴³Table 6 delineates the impact of these scavengers on the removal efficiency of TC antibiotics. Catalase, L-histidine, KI, EtOH, and TBA manifest conspicuous inhibitory effects on TC degradation. Intriguingly, BQ, conventionally linked with the superoxide radical (˙O₂⁻), exerts no substantial hindrance on the breakdown of TC, as discerned from the data in Table 6. The study proposes a chronological sequence for ROS involvement in TC degradation: H₂O₂ > ¹O₂ > ˙OH_surface > ˙OH_free. Furthermore, the validation of ¹O₂ and ˙OH participation is substantiated by electron paramagnetic resonance (EPR) spectra deploying TEMP and DMPO as the spin-trapping agents (Fig. S4†), respectively. These spectra underscore the formation of the TEMP complex with ¹O₂ in ethyl acetate and the DMPO complex with ˙OH in deionized water, underscoring the predominant role of ¹O₂ over ˙OH in TC degradation. Although numerical quantification via probe-based assays is beyond the scope of this study, the consistent suppression patterns in scavenger assays and dominant EPR signals support a qualitative hierarchy of ROS contributions, indicating ¹O₂ as the primary species, followed by H₂O₂-mediated ˙OH species. This conclusion is aligned with prior literature and validated through dual-mode detection (scavenging study and EPR signal) strategies.

Table 6 Degradation of TC in the presence of the 9AT catalyst, inhibited by the presence of scavengers (L-histidine, catalase, KI, EtOH, TBA, and BQ) and in an N₂ environment (reaction conditions: initial TC concentration = 20 mg L⁻¹, 9AT = 1.2 g L⁻¹, initial pH = 7.1, time = 120 min, temperature = 30 °C, agitation = 150 rpm)

Scavengers	Removal efficiency (%)	Scavenger for
No scavenger	96.5	—
L-Histidine	28.2	^1O2
Catalase	23.4	H2O2
KI	52.8	˙OHsurface
EtOH	63.5	˙OH
TBA	78.6	˙OHfree
BQ	94.4	˙O2^−
N2	9.3	O2

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3.5. Effect of air and nitrogen in the reaction medium

The emergence of ROS, which includes non-radicals such as ¹O₂ and radical ˙OH, originates from molecular oxygen (O₂) acquiring electrons from Mo⁺⁵ within the 9AT composite.^15,42 Section 3.9 extensively explores the intricacies of the reaction mechanism. To find out the source of oxygen in singlet oxygen (¹O₂), an examination of the impact of air and nitrogen (N₂) on the reaction system was undertaken (Table 6). The initial concentration of dissolved oxygen ([O₂]₀) in the reaction medium is 7.1 mg L⁻¹, with no aeration during the reaction. The introduction of N₂ into the reaction medium, resulting in a dissolved oxygen concentration below detection limit, led to a significant decrease in the degradation rate of TC from 96.5% to 9.3% within 120 min. This indicates that the atmospheric air facilitates the degradation of TC, while the absence of dissolved oxygen in the reaction system decelerates the degradation process. These suggest that the 9AT composite acts as a heterogeneous catalyst, with atmospheric oxygen (O₂) playing a crucial role as an oxidant in the degradation of TC. TiO₂ was used here as a coating material to protect the APM particles under near-neutral (pH ∼ 7) conditions.

3.6. Interference study

Extensive research has delved into the intricate breakdown of TC through oxidation, focusing on its dynamic interaction with prevalent inorganic anions and metal ions commonly present in wastewater. The investigation was further extended to see whether these metal ions and inorganic anions impede the efficacy of the AOP, creating a bottleneck by trapping ROS and thereby inhibiting the catalytic action.

In the realm of natural water, the influence of diverse inorganic anions significantly shapes the degradation process, ushering forth myriad effects such as radical quenching, the formation of less active species, and intricate interactions with the catalyst surface.^44–47 To embark on a comprehensive exploration of catalytic TC degradation, a series of detailed investigations were carried out on TC degradation with the 9AT catalyst. This encompassed a spectrum of inorganic anions, including Cl⁻, CO₃⁻², HCO₃⁻, H₂PO₄⁻, NO₃⁻, and SO₄⁻² with concentrations ranging from 0 to 100 mg L⁻¹. The culmination of these studies is assiduously compiled in Table 7. Notably, the introduction of inorganic anions, administered in the form of their respective sodium salts, reduced the degradation efficiency. Throughout the experiment, the initial pH of the solution was vigilantly maintained at 7.1. The results unveiled sulfate (SO₄⁻²) as the most formidable inhibitor among the tested inorganic anions, closely trailed by H₂PO₄⁻, NO₃⁻, CO₃⁻², Cl⁻, and HCO₃⁻. This observed impact is attributed to the innate ability of inorganic anions to either consume or scavenge reactive oxidizing radicals, consequently generating less reactive oxidative species during the TC degradation process.^44–47

Table 7 TC degradation efficiency in the presence of different inorganic anions and metal ions

Ions	Concentration (mg L^−1)	Removal efficiency (%)
Reaction conditions: initial TC concentration = 20 mg L^−1, 9AT = 1.2 g L^−1, initial pH = 7.1, temperature = 30 °C, reaction time = 120 min, and agitation = 150 rpm.
Blank (no ions added)	—	96.5
HCO3^−	20	95.2
	100	92.4
CO3^−2	20	90.2
	100	81.1
Cl^−	20	93.3
	100	86.7
H2PO4^−	20	82.6
	100	75.8
NO3^−	20	87.3
	100	78.5
SO4^−2	20	84.8
	100	71.2
Cu^+2	2	90.7
	10	83.6
Iron	2	92.5
	10	89.3
Mg^+2	2	95.8
	10	93.6
Mn^+2	2	94.1
	10	91.7

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Moreover, the experiments spanned a concentration spectrum from 0 to 10 mg L⁻¹ for metal ions, aiming to scrutinize the influence of prevalent interfering metal ions such as Cu⁺², Fe⁺²/Fe⁺³, Mg⁺², and Mn⁺² on the degradation process. Interestingly, the results revealed a conspicuous reduction in the efficiency of TC degradation in the presence of these dissolved metal ions, introduced in the form of their sulfate salts. Table 7 illustrates a discernible interference pattern, progressively intensifying with rising concentrations of these metal ions, with a disruption hierarchy of Cu⁺² > Fe⁺²/Fe⁺³ > Mn⁺² > Mg⁺². This disruption is partly attributed to the development of a slight positive charge on the surface of the microporous 9AT catalyst following the adsorption of these metal ions, corroborated by prior studies.^44,48,49 Additionally, Fig. S3† portrays the varying states assumed by the TC molecule depending on the solution pH. At a pH of 3.3, it adopts a cationic species (TC⁺), transitioning into a zwitterionic state (TC^±) within the pH range of 3.3 to 7.7 and shifting to an anionic state (TC⁻, TC⁻²) at a pH exceeding 7.7.¹⁵ The TC molecules take on a zwitterionic state, while the 9AT catalyst assumes an anionic state at a pH of 7.1. However, the slight positive charge acquired by the 9AT surface due to the adsorption of cationic metal ions hampers the effective adsorption of TC molecules on its surface, resulting in diminished degradation efficiency.

3.7. TC degradation in real wastewater matrices

Assessing the efficacy of 9AT catalysts within real water matrices is pivotal for their prospective implementation. Consequently, experiments were conducted to evaluate the degradation of TC in wastewater (refer to Fig. 9a and b for sample locations), and the outcomes are depicted in Fig. 9c. The characteristics of the water samples are succinctly summarized in Table S1.† The experimental setup involved spiking TC at a concentration of 20 mg L⁻¹ into collected wastewater. Fig. 9c and Table 5 reveal that when subjected to a 120 min reaction with the 9AT catalyst, TC degradation in pond water reached 90.2%, with a corresponding k value of 0.017 min⁻¹. In the case of natural municipal wastewater, TC experienced an 87.3% degradation at 120 min, accompanied by a k value of 0.016 min⁻¹.

Fig. 9 Collection point of (a) pond water and (b) real municipal wastewater, and (c) TC degradation performance of the 9AT catalyst in pond water and real wastewater (reaction conditions: initial TC concentration = 20 mg L⁻¹, 9AT = 1.2 g L⁻¹, initial pH = 7.1, temperature = 30 °C, agitation = 150 rpm).

Furthermore, the investigation extends to using a commercial TC tablet (provided by Abbott India Ltd.) in degradation studies. Fig. 9c illustrates the efficiency of the 9AT catalyst in dismantling the commercial TC tablet under optimal conditions, achieving a remarkable degradation of 97.6%. Notably, the performance of natural wastewater lags demonstrates a degradation of 87.3%, in contrast to pond water (90.2%) and distilled water (96.5%). The potential influence of common inorganic anions and metal ions on this diminished performance is acknowledged, as prior reports indicate a decline in the degradation efficiency with the increase in the concentration of these elements in natural water matrices.^44–49 The decrease in catalytic efficiency in real water matrices is consistent with the interference study (section 3.6), where a quantitative inverse relationship between specific ion concentrations (particularly SO₄⁻², H₂PO₄⁻, and Cu⁺²) and TC degradation is established. These ions are likely to inhibit the degradation process by scavenging reactive species and by altering the catalyst surface charge, thus reducing active site accessibility.

3.8. Changes in COD and TOC during TC degradation

This inquiry delves into a comprehensive examination of the degradation of TC through recording the time-dependent UV-visible absorption spectra, as illustrated in Fig. S5.† The investigation encompasses the continuous monitoring of COD and TOC removal over time to gauge the mineralization of TC. Notably, during 120 min oxidative reaction of TC in the presence of the 9AT catalyst under dark aerobic conditions, substantial removal rates were achieved, with COD and TOC exhibiting elimination rates of up to 94.3% and 71.4%, respectively. However, a captivating revelation from the analysis of time-dependent UV-visible spectra indicates a heightened elimination rate of 96.5% for TC under optimal conditions. This discrepancy suggests the existence of specific intermediates generated during the degradation process in the reaction mixture even after 120 min of reaction time.

3.9. Degradation mechanism

Leveraging the oxidative potential of oxygen and the catalytic efficacy of 9AT, ROS emerged in the course of the oxidative decomposition of TC. The intricate series of reactions, detailed in eqn (3)–(8),^{42,43,50–52} unfolds a chemical narrative. In this orchestrated ballet, the yellow Mo⁺⁶-infused 9AT catalyst metamorphosed into a blue Mo⁺⁵-laden counterpart as TC underwent oxidative breakdown (eqn (3)). Subsequently, when encountering dissolved O₂ and in situ produced H₂O₂, the blue Mo⁺⁵-loaded 9AT catalyst experienced oxidation, oscillating back to the yellow Mo⁺⁶ variant, as elegantly described in eqn (4), (5) and (7) (E⁰(O₂/H₂O₂) = 0.695 V & E⁰ (H₂O₂/˙OH) = 0.8 V). This cyclic inheritance of the 9AT catalyst becomes the linchpin, orchestrating a dynamic Mo-redox (E⁰(Mo⁺⁶/Mo⁺⁵) = 0.17 V) cycle pivotal for the degradation of TC. Simultaneously, this mechanism triggered a cascading sequence of free radical reactions (eqn (4)–(7)). The resulting ROS, encompassing hydroxyl radicals and singlet oxygen, intricately participated in the oxidative dismantling of TC owing to their higher reduction potential (e.g.¹O₂/O₂: 0.98 eV; ˙OH/OH⁻: 2.3 V) concerning Mo⁺⁶/Mo⁺⁵ reduction potential (0.17 V). This initiated the genesis of specific intermediates, culminating in their eventual mineralization. The plausible mechanism dictating the degradation of TC is as follows:

TC + [9AT(Mo⁺⁶)] → Oxidized products of TC + [9AT(Mo⁺⁵)] + H⁺ (3)

2[9AT(Mo⁺⁵)] + O₂ + 2H⁺ → 2[9AT(Mo⁺⁶)] + H₂O₂ (4)

[9AT(Mo⁺⁵)] + H₂O₂ → [9AT(Mo⁺⁶)] + ˙OH + OH⁻ (5)

Finally, eqn (4) and (5) can be combined as follows:

3[9AT(Mo⁺⁵)] + O₂ + 2H⁺ → 3[9AT(Mo⁺⁶)] + ˙OH + OH⁻ (6)

O₂ + [9AT(Mo⁺⁵)] + ˙OH → [9AT(Mo⁺⁶)] + ¹O₂ + OH⁻ (7)

TC + Reactive species (¹O₂, ˙OH) → Degraded products → CO₂ + H₂O (8)

3.10. Transformation products of TC degradation

In exploring transformation products (TPs) and the potential degradation pathway with the 9AT catalytic process, distinct samples were systematically collected during TC degradation at various time intervals (10, 40, and 90 min). Subsequent analysis through MALDI-ToF/ToF unveiled a spectrum of sixteen TPs, their molecular masses, and corresponding MS/MS spectra. The chemical structures of these TPs are elucidated in Fig. 10. By drawing on insights from existing research on TPs of TC, a proposed degradation mechanism is articulated in Fig. 11. This conceptualization posits that the TC molecules undergo non-selective attacks at various functional groups and sites, driven by an oxidative destruction mechanism rooted in the generation of reactive radicals. Consequently, each molecule exhibits multiple breakdown pathways.

Fig. 10 MS/MS spectra of transformation products (TP) of TC at (a and b) 10, (c) 40, and (d) 90 min reaction obtained using MALDI-ToF/ToF.

Fig. 11 Plausible TC degradation pathways of the 9AT catalyst.

Leveraging findings from prior studies,^{15,30,53–58} several mechanisms – hydroxylation, demethylation, deamination, decarbonylation, dehydroxylation, aromatic ring-opening reactions, and oxidation – contribute to TC degradation under the auspices of the 9AT catalyst under dark aerobic conditions. Fig. 11 further illustrates the interplay of these mechanisms, suggesting three potential degradation pathways. These routes originate from the capacity of active species including ¹O₂, ˙OH_surface, and ˙OH_free to attack TC molecules, ultimately mineralizing them into low-weight TPs. The consequential efficiency in TOC removal (71.4% in 120 min) underscores the potential mineralization of all generated molecules from these pathways into CO₂, H₂O, N-minerals, and more minor molecular constituents. While TPs are successfully characterized using MALDI-ToF/ToF, further structural confirmation and toxicity evaluation using LC-MS/MS and HPLC methods remain an important future direction to fully validate the proposed degradation pathways and ensure environmental safety of the intermediates. Although TOC and COD reductions suggest substantial mineralization, the potential ecotoxicity of intermediate TPs remains an important concern. While initial structural analysis indicates lower toxicity than the parent TC molecule,^54,56,58 future studies may incorporate QSAR modeling (e.g., ECOSAR, VEGA) and bioassays to validate the ecological safety of these intermediates.

3.11. Leaching of molybdenum from 9AT during TC degradation

Investigating the leaching tendencies of molybdenum ions from the 9AT catalyst is a pivotal scientific quest. The behavior of these metal ions was evaluated after a 120 min reaction, revealing a molybdenum release of 1.37 mg L⁻¹ into the solution phase. This concentration remains well below the industrial wastewater discharge limit of 40 mg L⁻¹ (ref. 59) and a local regulatory threshold of 4.32 mg L⁻¹, as established by the City of Salem, Oregon, USA, in July 2007.⁶⁰ These findings highlight the stability of the 9AT composite and reinforce its potential as a sustainable catalyst for the efficient mineralization of TC antibiotics.

3.12. Comparative account on TC degradation using other composites

In exploring AOPs designed to degrade TC using various POM composites, a comprehensive comparative assessment was conducted on catalysts employed in these processes. This scrutiny specifically targeted the efficacy comparison with the 9AT composite. The comparative analysis considered multiple factors including experimental conditions and performance metrics, as summarized in Table 8, drawing from relevant prior studies.^17–29 The results clearly indicate that the 9AT composite exhibits degradation efficiency and catalytic stability comparable to those of previously reported POM-based composite catalysts. Unlike the conventional POM–TiO₂ hybrids that require strongly acidic media for catalytic function,^28,29 the 9AT composite exhibits high oxidative performance at neutral pH (7.1), achieving over 96% TC degradation under ambient and dark conditions. This pH-tolerance and high efficiency is a result of the synergistic interaction between microporous APM and the TiO₂ coating, offering a structurally stable and redox-efficient platform for practical wastewater remediation under ambient conditions.

Table 8 Summary of tetracycline degradation studies on various POM composites

AOP	Catalyst/reagents	Experimental conditions	Removal efficiency	References
Polyoxometalate composites	Artificial chloroplast-phosphotungstic acid–iron oxide microbox heterojunctions penetrated by carbon nanotubes (HPWx@Fe2O3-CNTs)	[TC]0 = 10 mg L^−1; HPWx@Fe2O3-CNTs = 0.25 g L^−1; 300 W xenon light irradiation of 420 nm @ 300 mW cm^−2	100% in 100 min	Sun et al. 2022 (ref. 17)
Polyoxometalate composites	Vanadium substituted polyoxometalate doped magnetic carbon nitride (CoWV/mCN)	[TC]0 = 10 mg L^−1; CoWV/mCN = 0.5 g L^−1; 300 W xenon light irradiation of 420 nm	82.4% in 120 min	Zhang et al. 2021 (ref. 18)
Polyoxometalate composites	Polyoxometalates supported on biochar (CoPMoV/C)	[TC]0 = 20 mg L^−1; CoPMoV/C = 0.15 g L^−1; peroxymonosulfate (PMS) = 0.17 mM; 300 W xenon light irradiation of 420 nm	100% in 30 min	Zhu et al. 2023 (ref. 19)
Polyoxometalate composites	Multivalent supramolecular self-assembly between β-cyclodextrin derivatives and polyoxometalate (POM/EDA-CD)	[TC]0 = 100 μL of 1 mM solution; 0.055 mM POM/0.03 mM EDA-CD = 3 mL; [H2O2]0 = 50 μL (30% w/w); 50 W white mercury light irradiation	∼100% in 25 min	Wang et al. 2019 (ref. 20)
Polyoxometalate composites	Polymer composite of phosphotungstic acid and polyaniline (PWA/PAN)	[TC]0 = 25 μM; PWA/PAN = 0.5 g L^−1; [H2O2]0 = 50 mM (30% w/w)	100% in 30 min	Liu et al. 2022 (ref. 21)
Polyoxometalate composites	Nanospherical α-Fe2O3 supported on 12-tungstosilicic acid (α-Fe2O3/12-TSA·7H2O)	[TC]0 = 30 mg L^−1; α-Fe2O3/12-TSA·7H2O = 0.15 g L^−1; [H2O2]0 = 0.1 g l^−1 (30% w/w); 15 W UV light irradiation of 254 nm	97.39% in 50 min	Saghi and Mahanpoor 2017 (ref. 22)
Polyoxometalate composites	Fe-polyoxometalates nanoparticles on porous and ultrathin g-C3N4 nanosheets with nitrogen vacancies (Fe-POM/CNNS-Nvac)	[TC]0 = 20 mg L^−1; 45-Fe-POM/CNNS-Nvac = 1.00 g L^−1; [H2O2]0 = 10 mM (30% w/w); 300 W xenon light irradiation of 420 nm	96.5% in 18 min; 78.5% TOC removal in 18 min	Jiang et al. 2020 (ref. 23)
Polyoxometalate composites	Polyoxometalate intercalated La-doped NiFe-LDH (NiFeLa{HPW}-LDH)	[TC]0 = 5 mg L^−1; NiFeLa{HPW}-LDH = 0.10 g L^−1; peroxymonosulfate (PMS) = 2.00 mM	95.4% in 60 min	Wang et al. 2021 (ref. 24)
Polyoxometalate composites	Polyoxometalate anchored zinc oxide nanocomposite (ZnO@PDA/Cu-POMs)	[TC]0 = 50 mg L^−1; ZnO@PDA/Cu-POMs = 50.00 g L^−1; 300 W xenon light irradiation of 420 nm	90.75% in 90 min	Xing et al. 2023 (ref. 25)
Polyoxometalate composites	Flower-ball-like ZnIn2S4@hollow dodecahedral polyoxometalate (K3PW12O40)@flower-shell-like ZnIn2S4/Ag2S yolk@shell (ZIS@HD-KPW@ZIS/AS)	[TC]0 = 20 mg L^−1; ZIS@HD-KPW@ZIS/AS = 0.20 g L^−1; 300 W xenon light irradiation of 420 nm	∼99% in 120 min	Wu et al. 2022 (ref. 26)
Polyoxometalate composites	Au nanoparticle-decorated polyoxometalate/zeolitic imidazolate framework-8 nanostructure (ZIF-8@PTA@AuNP)	[TC]0 = 10 mg L^−1; ZIF-8@PTA@AuNP = 0.5 g L^−1; 125 W mercury UV light irradiation	97.0% in 60 min	Beni et al. 2020 (ref. 27)
Polyoxometalate composites	K5CoW12O40/TiO2 composite	[TC]0 = 50 mg L^−1; 30-K5CoW12O40/TiO2 = 5.00 g L^−1; pH = 5; 18 W fluorescent light irradiation	∼100% in 120 min	Mahmoodi et al. 2022 (ref. 28)
Polyoxometalate composites	Mesoporous TiO2-doped with polyoxometalates [PMo10V2O40]^5− (PMV) and loaded with Ag nanoparticle composite (PMV-TiO2/Ag0.1)	[TC]0 = 20 mg L^−1; PMV-TiO2/Ag0.1 = 1.00 g L^−1; pH = 6; 300 W xenon light irradiation of 420 nm	90.0% in 80 min	Zhao et al. 2021 (ref. 29)
Polyoxometalate composites	Ammonium phosphomolybdate–TiO2 composite (AT)	[TC]0 = 20 mg L^−1; [9AT] = 1.2 g L^−1; pH = 7.1	96.5% in 120 min; 71.4% TOC removal in 120 min	This study

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4. Conclusions

A novel method has been developed for the synthesis of an ammonium phosphomolybdate–titanium dioxide (AT) composite material, which has been successfully employed as a catalyst for the degradation of the TC antibiotic. Among the assorted composites synthesized, the 9AT (APM : TiO₂ = 9 : 1 w/w) variant demonstrated superior performance in TC degradation under ambient aerobic dark conditions at near-neutral pH. Here TiO₂ acted as a coating/stabilizing material for APM. Thus, the dissolution of APM could be restricted during TC degradation reaction. Remarkably, the 9AT composite exhibited a tremendous capability to degrade 96.5% of TC (20 mg L⁻¹) within 120 min at a neutral pH of 7.1, showcasing a degradation rate coefficient of 0.024 min⁻¹. Moreover, the examination of TC degradation efficiency in authentic water matrices such as pond water and natural wastewater revealed a decline, which was attributed to the interference of dissolved inorganic anions and metal ions in natural water samples. The 9AT composite demonstrated robust stability even after eight cycles of TC degradation. Radical scavenging experiments elucidated that in situ-produced H₂O₂, ¹O₂, ˙OH_surface, and ˙OH_free are the primary reactive species, with ˙OH playing a comparatively lesser role in the degradation process. Lastly, a plausible three-stage degradation pathway was proposed based on the findings from the mass spectra of MALDI-ToF/ToF. This study catalyzes further exploration into the unique degradation potential of the 9AT composite against a spectrum of pollutants.

Data availability

All relevant data are available within the manuscript.

Author contributions

Debasish Pal: writing – original draft, visualization, methodology, and investigation. Anjali Pal: writing – review & editing, supervision, and conceptualization.

Conflicts of interest

The authors affirm that they have no known financial or personal affiliations that would have appeared to conflict with the work described in this study.

Acknowledgements

The authors are grateful to IIT Kharagpur and Ministry of Education, Government of India for providing the instrumental facility and financial support.

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Footnote

† Electronic supplementary information (ESI) available. See DOI: https://doi.org/10.1039/d5cy00234f

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