1 Introduction

Water sources are increasingly contaminated with organic and inorganic pollutants. About 1.1 billion people are in risk contracting waterborne diseases due to poor water access [1]. Chromium (Cr(VI)) and tetracycline (TCL) are two common inorganic and organic pollutants that can contaminate water sources and posing a significant environmental and health risks [2, 3]. Chromium is an abundant element in the Earth's crust. However, due to its hexavalent state (Cr(VI)), which is highly toxic and a known carcinogen becomes a potent water contaminant [4]. Cr(VI) pollution typically arises from industrial processes, such as electroplating, leather tanning, and stainless steel production [5]. Ingesting or coming into contact with Cr(VI)-contaminated water can lead to severe health issues, including lung cancer and skin problems [6]. Furthermore, Cr(VI) could significantly harm the aquatic life and disrupt ecosystems. Interestingly, chromium in its trivalent state (Cr(III)) has less toxic effects, which can also be readily removed or adsorbed on different materials, including photocatalysts [7]. Although there are strong restriction on Cr(VI) permissible levels in drinking water and industrial wastewater, it still has higher levels of discharge to mitigate its adverse effects which is needed to be urgently addressed [8].

As a broad spectrum antibiotic, TCL has been using for various infections in human and veterinary medicine [9], and the untreated wastewater from pharmaceutical industries and the agricultural runoff led their presence in many water sources [10, 11]. Since TCL is chemically stable, it is challenging to eliminate it under natural conditions [12]. Additionally, higher usage of these TCL antibiotics can lead to various health issues, including gastroenteritis, hepatotoxicity, chronic poisoning, and some allergic reactions [13]. Therefore, eradicating these inorganic and organic pollutants from the aqueous environment is important and essential [14].

Traditional techniques, such as chemical precipitation, adsorption, ion exchange, and biological treatment, have removed Cr(VI) and tetracycline pollutants from water resources [15]. The primitive technological approaches such as coagulation, adsorption, ultrafiltration, air stripping, flocculation, and precipitation were evaluated for their efficiency and limitations in wastewater treatment [16, 17]. Notably, these non-destructive methods result only in the transformation of drug molecules into solids or sludge, which, in turn, produce secondary pollutants that needs to be treated and purified. Also, some of these techniques may not be specific to the target pollutant, leading to remove other beneficial ions or molecules from the water [18]. Given these limitations, a growing need exists to develop and implement more efficient, cost-effective, and environmentally friendly advanced technologies for removing Cr(VI) and TCL pollutants from water resources [19, 20]. Advanced methods like membrane filtration, electrochemical treatment, and nanomaterial-based approaches are being explored as alternatives to address these challenges [21]. Using methods like membrane filtration, electrochemical treatment, and nanomaterial-based approaches, the pollutants can interact at a molecular level, resulting in more efficient and effective removal from water. This technology can be beneficial for both industrial and municipal wastewater treatment. However, all these above-mentioned methods involved the usage of overpriced and unstable oxidants that also generate unwanted secondary pollution during the degradation process. Thus, it is an urgent need to design a suitable and efficient technology to remove inorganic and pharmaceutical pollutants from the environment.

In particular, heterogeneous photocatalysis is adaptable for both point-of-use and large-scale water treatment applications. It is a versatile solution for removing Cr(VI) and TCL pollutants from various water bodies [22]. This process leverages the properties of specific photocatalysts activated by light to drive chemical reactions that lead to pollutant reduction and degradation [23]. However, effective utilization of abundant sunlight by constructing heterojunction photocatalysts with enhanced photogenerated charge separation rate is important in the degradation process [24, 25].

Cobalt titanate (CoTiO3) belongs to the ABO3 type perovskite oxide, which has a good ability for visible-light absorption, faster mobility of charge carriers, and good photostability. It has attracted substantial attention in high-κ gate dielectric ceramics, gas sensors, and pigments [26, 27]. Regardless of its suitable energy band gap (2.25 eV), CoTiO3 suffers from its poor efficiency during photocatalysis. However, the reduction in the recombination rate of photogenerated carriers of CoTiO3 with full utilization of sunlight can be achieved by constructing a heterojunction between CoTiO3 (p-type) and another semiconducting material [26, 28]. Based on this concept, Yang et al. proposed a core–shell CoTiO3@MnO2 heterostructure as a visible-light-induced photocatalyst for removing TCL, methyl orange, metronidazole, and bisphenol pollutants in wastewater. This composite possesses 3 ~ 4 times increased efficiency than the pure CoTiO3 and MnO2 [29]. Similarly, Biyun et al. proposed a C-TiO2/CoTiO3 type-II heterojunction with a core/shell structure that was successfully utilized to remove 99.6% of the antibiotic enrofloxacin under visible light illumination over 120 min [30]. In addition to this, the use of CoTiO3/g-C3N4, CoTiO3/Ti3C2TX MXene, and CoTiO3/UiO-66-NH2 based materials are also reported [26, 31, 32]. Although few studies were conducted on constructing CoTiO3-based heterojunction materials for various environmental applications, the outcomes are not well intended [33]. This problem can be overcome by introducing an interface between two semiconductors with a suitable band alignment to improve the spatial separation of charge carriers, thus facilitating reduction/oxidation reactions between the organic/inorganic contaminants and the charge carriers [34].

Recently, bismuth vanadate (BiVO4) has garnered considerable attention as a photo-catalyst due to its numerous advantages, including its low bandgap, nontoxicity, cost-effective, and photo-corrosion resistance with efficient visible light activity [35]. In addition, BiVO4 has also been reported to be an antibacterial material. However, several drawbacks including low specific surface area and high recombination rate of electron–hole limits its application [36]. To enhance the photocatalytic function of BiVO4 (n-type), a variety of BiVO4‐based composites has been prepared, including BiVO4/g-C3N4, BiVO4/Ag2O, Fe2O3/BiVO4, r-GO/BiVO4, and BiVO4/TiO2 [37,38,39,40,41]. These photocatalysts typically have low adsorption capacities for organic and inorganic pollutants, hindering their commercialization [42]. As a result, it is imperative to develop novel composites with enhanced adsorption capacities.

Metal–organic frameworks (MOFs) are porous, crystalline materials of metal ions/clusters coordinated to organic ligands [43]. Metal–organic frameworks (MOFs) have recently been identified as a class of supramolecular non-noble materials suitable for photocatalytic degradation and reduction reactions [44,45,46]. Compared to conventional materials, MOFs have a porous structure, a large specific surface area, and the ability to be tailored. Furthermore, MOF-based materials can effectively separate photogenerated carriers, making them promising materials for photocatalysis [47]. Several composite materials have been synthesized as photocatalysts, including Cu2O/N-CQD/ZIF-8, UiO-66/Bi2WO6, MIL-125(Ti)/BiOCl, and MIL-53(Fe)/Sn3O4 [48,49,50,51]. Moreover, they all displayed better photocatalytic activity than their pure counterparts. Most of these composites contain semiconductors as the primary photocatalyst, while MOFs serve as a support or co-catalyst [52]. Due to their abundance in nature, Fe-based MOFs have attracted attention due to their low price; in addition to its excellent ability to absorb visible light, MIL-53 (Fe) also displays remarkable physicochemical properties that make it an ideal framework and co-catalyst material [53]. For instance, Yan et al. found that MIL-53(Fe) can efficiently remove dyes and photoreduce Cr(VI) under visible light [54]. Recently, Liu et al. applied MIL-53(Fe) to remove ibuprofen and demonstrated its practical effect in visible light [55]. Despite the good photocatalytic capacity of the pristine MIL-53(Fe), the high rate of carrier recombination prevents it from being used in practical applications.

Based on the above-mentioned points, in this study a hierarchical CoTiO3/BiVO4 (p-n heterojunction) nanorods embedded in MIL-53 (Fe) were constructed, with excellent visible light harvesting capabilities and abundant reaction sites for the photoreduction of Cr(VI) and degradation of TCL. Among the various prepared catalysts, the CT/BV@Fe-MOF heterostructure displayed improved performance. The influence of pH, catalyst loading, and Cr(VI) and drug concentrations was further evaluated in detail. The radical quenching experiments and LC–ESI–MS analysis were carried out to determine the degradation pathway of TCL. Furthermore, the TEST program was used to predict the toxicity of the products generated during the degradation. The CT/BV@Fe-MOF nano-photocatalyst shows the promising solution to address the complex environmental challenges and underscores the potential of advanced materials in environmental remediation.

2 Materials and methods

2.1 Chemicals and instrumentations

For the fabrication of CT/BV@Fe-MOF heterostructure, chemicals such as Cobalt acetate (Co(OAc)2.4H2O, 99.9%), Titanium butoxide (99.0%), Polyvinyl pyrrolidone (PVP, average Mw 58000), Ethylene glycol (99.0%), Bismuth (III) nitrate pentahydrate (Bi(NO3)3.5H2O, 98.0%), Ammonium metavanadate (NH4VO3, 99.5%), Nitric acid (HNO3, 65–68%), Ferric chloride hexahydrate (FeCl3·6H2O, 98.0%), Benzenedicarboxylic acid (> 98%), and N, N-dimethylformamide (DMF, > 99.8%) were procured from Aladdin chemicals, Shanghai, China. For photocatalysis studies, The Chromium source (K2Cr2O7, 99.0%), 1,5-diphenylcarbazide (98.0%), were purchased from Nanjing Chemical Reagent Co., Ltd. Tetracycline (98.0 − 102.0%) were procured from Sigma Aldrich, Shanghai, China). For characterizing the prepared catalysts, the following techniques were used such as p-XRD (BRUKER D8 Advance), FE-SEM (ZEISS Sigma 300), TEM-SAED (ThermoFisher Talos F200S G2 electron microscopes), UV–Vis-DRS (Shimadzu UV-360), and XPS (ThermoScientific K-AlphaX).

2.2 Synthesis of CoTiO3 nanorods

In a typical synthesis, 10.0 mmol of Co(OAc)2.4H2O and 1.0 g PVP was added in a 70 mL of ethylene glycol solution while stirring. Up on dissolution of the above mixture, 8.0 mmol of titanium butoxide was added dropwise and stirred for another 12 h. The product was then washed with ethanol and dried in a vacuum oven at 65 °C overnight. After obtaining the light pink solid powder, it was placed in a covered crucible and calcined at 700 °C for 2 h at a rate of 5 °C/min to obtain a green powder denoted as CoTiO3.

2.3 Preparation of CoTiO3/BiVO4 nanocomposite

The CoTiO3/BiVO4 composite was prepared by a hydrothermal method. Firstly, 5.0 mmol of Bi(NO3)3.5H2O was added to 15 mL of HNO3 (10%) followed by addition of 0.2 g of the prepared CoTiO3 and dispersed through ultrasonication for 30 min (solution-A). Meanwhile, in a separate beaker, 5.0 mmol of NH4VO3 was added to 50 mL of distilled water at 75 °C and stirred until to get a clear solution (solution-B). Solution-B was slowly added to solution-A, and the pH of the reaction mixture was maintained at pH-7.0 using ammonia solution. The solution was then stirred for another 1.0 h before being put into a Teflon-lined stainless-steel autoclave for 24 h at 180 °C. After cooling, the final product was washed with DI water and ethanol and dried overnight at 80 °C. The pristine BiVO4 was prepared by following the similar procedure without addition of CoTiO3.

2.4 Synthesis of CT/BV@Fe-MOF heterostructures

In the first step, 26 ml of DMF solution was added to 5.0 mmol of FeCl3.6H2O, to which 5.0 mmol of 1,4-benzene dicarboxylic acid (1,4-BDC) was slowly added and stirred for about 30 min. A homogeneous yellowish solution was obtained by adding 2.0 mL of HF (40 wt%) dropwise to the above mixture while constantly stirring. In this homogeneous solution, 0.2 g of CoTiO3/BiVO4 nanocomposite was added and stirred at room temperature for about 4.0 h. It was then transferred to a Teflon-lined autoclave and heated to 160 °C for 22 h. After bringing the autoclave to normal, the final product was washed several times with ethanol and water. The separated product was dried overnight at 65 °C, while the pure MIL-53(Fe) nanorods were synthesized similarly without adding nanocomposite.

2.5 Photocatalytic activity

The degradation of TCL and the reduction of Cr(VI) were carried out to test the photocatalytic performance of the synthesized photocatalysts. A 300 W Xe lamp was used as the light source for all the experiments with a filter (≤ 420 nm). In the TCL degradation experiments, the adsorption–desorption isotherm was achieved by adding 50 mg of the photocatalyst in 100 mL of TCL solution (30 ppm), and stirred under dark for 15 min. Then the visible light was turned on and 3.0 mL aliquots were collected at regular intervals, followed by centrifugation and filtration (0.2 μm filters) before carrying out HPLC analysis, in which acetonitrile (20% v/v) and water (80% v/v) as the mobile phase with 1.0 mL/min flow rate and a UV detector set at 354 nm. A similar procedure was followed to determine the degradation products in which the HPLC was connected to a tandem mass spectrophotometer (LC-ESI/MS). The reduction of Cr(VI) was accomplished using a solution of Cr(VI) (50 ppm) dispersed with 30 mg of the photocatalyst and K2Cr2O7 as a Cr(VI) source. The mixture was stirred continuously in the dark to reach adsorption equilibrium for 15 min. Finally, Cr (VI) content was determined using the diphenyl carbazide method (DPC).

3 Results and discussion

3.1 Characterization

The crystal structure, phase purity, and structural characteristics of pristine CoTiO3, BiVO4, MIL-53 (Fe), and CT/BV@Fe-MOF nanocomposites were identified by p-XRD analysis (Fig. S1). The pristine CoTiO3 annealed at 700 °C was identified with a distinct diffraction peak at 2θ values of 24.02°, 32.98°, 35.43°, 40.59°, 49.02°, 53.64°, 56.90°, 62.06°, 63.69° and 71.03°, which was assigned to the (hkl) planes of (012), (104), (110), (113), (024), (116), (018), (124), (300) and (119), corresponding to the facets of the crystalline rhombohedral CoTiO3 phase (JCPDS Card No: 15–0866). Similarly, the bare BiVO4 exhibits a diffraction peak at 15.28°, 18.81°, 19.08°, 28.87°, 30.77°, 34.57°, 35.39°, 40.01°, 42.72°, 46.80°, 47.34°, 50.33°, 53.32°, 58.49° and 59.57°, which related to (002), (101), (011), (121), (004), (200), (002), (211), (015), (213), (204), (220), (116), (312), and (206) planes, respectively (JCPDS Card No: 14–0688). The diffraction peaks obtained at 2θ = 9.38, 12.65, 17.65, and 25.73 corresponds to the planes of pristine MIL-53 (Fe) (CIF file no: 690314–690316). According to the XRD pattern of the CT/BV@Fe-MOF composite, characteristic peaks of CoTiO3, BiVO4, and MIL-53 (Fe) coexist on the surface, which indicates that a heterostructure has been formed between them. In addition, no other diffraction peaks were found, indicating higher material purity. Debye–Scherrer equation was used to calculate the crystalline size of the prepared materials, and the CT/BV@Fe-MOF composite had an average crystalline size of 64.57 nm.

UV–vis DRS spectroscopy was obtained to examine the light absorption properties and optical band gap of the prepared samples (Fig. S2a). The absorption edges of the synthesized BiVO4 appear at 514 nm, while the CoTiO3 exhibited absorption bands from UV to visible light in a wide spectral range at ~ 491 nm results from the charge transfer interaction between O−2 and Ti+4. Furthermore, two other absorption bands appeared for CoTiO3 at 538 and 609 nm, which were caused by the splitting of Co (3d8) orbital in the crystal field, implying the existence of Co+2. Similarly, the MIL-53 (Fe) exhibited light absorption ranges from 200 ~ 600 nm, in which the peak around 300–500 nm is ascribed to the spin-allowed d-d transition of Fe3+. Incorporating BiVO4/CoTiO3 nanocomposite into MIL-53 (Fe) surfaces enhanced its ability to capture the visible light, supporting its effectiveness as a visible light-active photocatalyst. Furthermore, based on absorption spectra, the band gap energies of samples were calculated using Tauc plot relations using the equation (αhν)2 = A (hν-eg)n. Where “hν” is the incident photon energy, “α” is represented as the absorption coefficient, “A” is the proportionality constant, and “Eg” denotes the band gap energy. However, the band gap energy of the prepared photocatalyst materials was determined by plotting (αhν)2 versus hν (Fig. S2b). Accordingly, the calculated bandgap energy values for BiVO4, CoTiO3, MIL-53 (Fe), and CT/BV@Fe-MOF composite samples are 2.65, 2.50, 2.30, and 2.17 eV, respectively.

The as-synthesized CoTiO3 exhibited a one-dimensional microrod-like structure composed of many tightly packed nanoparticles (Fig. 1a). On the other hand, FE-SEM images of bare BiVO4 shows a coral-like structure with spherically shaped microporous particles in an agglomerated state which appears like humps of numerous numbers of small noodles (Fig. 1b). Furthermore, the low-resolution FE-SEM images of pristine MIL-53 (Fe) demonstrate a rectangular microrod morphology with a uniform and smooth surface (Fig. 1c). For the FE-SEM images of CT/BV@Fe-MOF composite (Fig. 1d), it is notable that the nanocomposite particles of CoTiO3/BiVO4 are uniformly distributed on the surface of MIL-53(Fe) without affecting the primary crystal of the Fe-MOF. The obtained FE-SEM results are in agreement with the p-XRD results. Furthermore, these results demonstrated the successful fabrication of CT/BV@Fe-MOF heterojunction, which might enhance the photocatalytic activity by improving the separation of the photo-excited electron and hole charges.

Fig. 1
figure 1

FE-SEM images of a CoTiO3, b BiVO4, c Fe-MOF, and d CT/BV@Fe-MOF; e–f HR-TEM images and g-o HAADF elemental mapping of CT/BV@Fe-MOF nanocomposite

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HR-TEM images were obtained for tri-composite CT/BV@Fe-MOF nanohybrids and the images revealed that the CoTiO3 and BiVO4 are uniformly attached to the surface of MIL-53 (Fe), consistent with the FE-SEM results. The HR-TEM image (Fig. 1e-f) show lattice fringes at d-spacings of 0.25 nm, 0.29 nm, and 0.32 nm, correlating to crystal planes (110), (004), and (111) of CoTiO3, BiVO4, and MIL-53 (Fe). These evidences confirm the formation of heterojunctions by enhancing the ability to transport charge and increasing the efficiency of photocatalysis. In addition, the high-annular angle dark field (HAADF) STEM imaging along with the elemental mapping substantiate that in the CT/BV@Fe-MOF composite sample all the components (Co, Ti, Bi, V, Fe, O, and C) which are intact during heterojunction formation to promote the enhanced catalytic activity during photodegradation process (Fig. 1g-o).

The BET-specific surface area and BJH analysis of all the synthesized materials showed a type-IV isotherm pattern with a clear H3 hysteresis loop, designating a mesoporous structure (Fig. 2a, b). Based on the results, the pristine MIL-53 (Fe) exhibits a high surface area value of 209.73 m2g−1, a total pore volume of 0.184 cm3g−1, and an average pore diameter of 3.04 nm. On the other hand, the bare BiVO4 has the least surface area of 10.64 m2g−1 with a pore volume of 0.0106 cm3g−1 and an average pore diameter of 3.49 nm. Additionally, the BET surface area of CoTiO3 was measured to be 45.73 m2g−1 with pore volumes and diameters of 0.117 cm3g−1 and 3.40 nm, respectively. However, after forming the heterostructure, the composite material CT/BV@Fe-MOF exhibits an impressive surface area of around 105.62 m2g−1 (pore volumes of 0.094 cm3g−1 and pore diameters of 3.05 nm), which is significantly greater than the pure BiVO4 and CoTiO3. As a result of this increased surface area, more reaction sites and adsorption points could be provided on the nanocomposite surface, which facilitates the photocatalytic reduction and degradation of both inorganic and organic pollutants [42, 56].The detailed information on the surface area, pore volume, and pore size of all samples can be found in Table 1.

Fig. 2
figure 2

a N2 adsorption desorption isotherm (b) BJH pore size distribution spectra of CoTiO3, BiVO4, Fe-MOF, and CT/BV@Fe-MOF nanocomposite

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Table 1 The specific surface area, pore volume and pore radius of the prepared samples
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The XPS analysis was carried out to determine the oxidation state and elemental composition of the prepared nanocomposites. The results indicating the presence of Co 2p, Ti 2p, Bi 4f, V 2p, Fe 2p, C 1 s, and O1s elements in the CT/BV@Fe-MOF, confirming the photocatalyst's purity (Fig. S3a). A deconvoluted XPS peak for the Co 2p exhibits distinct binding energy at 780.5 and 796.3 eV, which suggests + 2 oxidation state of Co (Fig. S3b). Likewise, the deconvoluted Ti 2p spectrum (Fig. S3c) reveals two significant peaks at binding energies of 457.8 and 463.7, corresponding to Ti 2p3/2 and Ti 2p1/2 of Ti4+, respectively. Meanwhile, high-resolution Bi spectra (Fig. S3d) show two prominent peaks at 158.6 and 163.9 eV, corresponding to Bi 4f7/2 and Bi 4f5/2 of Bi, indicating that Bi3+ is presented in BiVO4. Similarly, the deconvoluted V5+ also displayed two distinct peaks at 516.2 and 523.8 eV, corresponding to V 2p3/2 and V 2p1/2 orbital states (Fig. S3e). The Fe 2p spectrum corresponding to MIL-53 (Fe) shows peaks at 724.3 and 710.4 eV associated with Fe 2p1/2 and Fe 2p3/2, respectively, with a difference of 13.9 eV, suggesting the presence of Fe3+ in Fe-oxo clusters. Furthermore, satellite peaks corresponding to the same were observed at 715.3 eV and 719.4 eV (Fig. S3f). The deconvoluted C 1 s spectrum corresponding to MIL-53 (Fe) exhibits three significant peaks at 288.1, 285.0, and 284.0 eV, which are corresponds to the C = O in the carboxylic group of terephthalates, the C-O between the benzene ring and Fe-oxo cluster, and the C = C in the benzene ring, respectively (Fig. S3g). Furthermore, the O 1 s characteristic peaks were deconvoluted into two peaks at 529.3 and 530.7 eV, corresponding to Co–O-Ti and Bi-O-V lattice oxygen. In addition, the peak induced by the oxygen atom in the carboxylic group corresponding to MIL-53 (Fe) appeared at a binding energy of 533.4 eV (Fig. S3h). Thus, these results suggested that all the elements are uniformly distributed in the as prepared CT/BV@Fe-MOF nanocomposite.

3.2 Photocatalytic Cr(VI) reduction

To verify the photocatalytic reduction performance of CT/BV@Fe-MOF nanocomposite, an aqueous Cr(VI) solution was used as a test solution. Photocatalytic reduction efficiency was calculated using the following Eq. (1).

$$\mathrm{Photocatalytic}\;\mathrm{reduction}\;\mathrm{efficiency}\;\left(\%\right)=\frac{\left(\mathrm{Co}-\mathrm{Ct}\right)}{\mathrm{Co}}\times100$$
(1)

where C0 represents the initial concentration of a Cr(VI) pollutant, and Ct represents the final concentration of a Cr(VI) pollutant after t-min of photocatalysis. As a result of 90 min of visible light illumination, it has been observed that the heterojunction composite obtained have a remarkable reduction efficiency when compared to other pristine materials, resulting in reduction percentages of 5.8%, 33.5%, 47.1%, 55.3%, 60.1% and 99.2% for BiVO4, CoTiO3, MIL-53 (Fe), CoTiO3/Fe-MOF, BiVO4/Fe-MOF and CT/BV@Fe-MOF, respectively (Fig. 3a). Due to the rapid recombination of photogenerated electrons and holes, bare materials have limited photocatalytic activity, while the CT/BV@Fe-MOF heterojunction with a nanorod structure improves the efficient transfer of photogenerated electrons and holes, which is a critical parameter for improved Cr(VI) reduction.

Fig. 3
figure 3

a Influence of photocatalytic reduction of Cr(VI) over BiVO4, CoTiO3, MIL-53 (Fe) and CT/BV@Fe-MOF nanocomposite; effect of (b) solution pH, c photocatalyst quantity, and d Cr(VI) concentration; e reusability study, f radical trapping experiments

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It is estimated that the original Cr(VI) solution has a pH of approximately 5.2 and is predominantly present in the form of HCrO4− and Cr2O72−. An investigation on the effect of pH during photocatalytic reduction of Cr(VI) was conducted at different pH levels (2.0–11.0) using 0.1 N HCl and 0.1 N NaOH. As a result (Fig. 3b), the nanocomposite of CT/BV@Fe-MOF exhibits a remarkable adsorption and reduction efficiency at pH 3.0 (97.5%) since the cationic charge of photocatalyst can electrostatically interact with negatively charged HCrO4− and Cr2O72− species. Consequently, Cr(VI) ions at this specific pH can efficiently bind to the photocatalyst and be reduced to Cr(III). In addition, the decline in photocatalysis rate at pH 2 is caused by the large quantity of H + ions generated in the reaction mixture in acidic media. These ions not only deposit on the surface of the photocatalyst but also partially block the pores in the material, thereby preventing Cr(VI) molecules from easy access to active sites. In contrast, above pH 3.0, the surface of the catalyst becomes negatively charged, resulting in a repulsive force between the material and Cr2O72− species. Additionally, the precipitation of Cr(OH)3 with an increase in pH will block the active sites of CT/BV@Fe-MOF, decreasing reduction efficiency.

To further determine how the concentration of photocatalyst affects the reduction of Cr(VI) during photocatalysis, various concentrations of CT/BV@Fe-MOF (25 mg to 150 mg) were used (Fig. 3c). The results indicated that the reduction efficacy increases as catalyst concentration increases from 25 to 75 mg, and thereafter the pattern decreases. This is attributed to the light scattering, and some saturation occurred as the solution opacity increased.

Figure 3d shows the effect of metal ion concentration on the photocatalytic reduction of Cr(VI). At first, optimal conditions (75 mg of catalyst, pH-3.0) were established to equilibrate Cr(VI) species with concentrations ranging from 10–70 ppm. Then, the reduction efficiency of Cr(VI) pollutants was evaluated over varying periods. Due to the availability of more active sites on the surface of the photocatalyst, a lower Cr(VI) concentration (10–30 ppm) resulted in a faster reduction rate. Conversely, when Cr(VI) concentrations increased above 50 ppm, the reduction efficiency of Cr(VI) decreased rapidly because a high amount of Cr(VI) species adsorb on the photocatalyst surface, which ultimately reduced the penetration of photons and consequently decreased the photo-generated radicals. As a result, the 50 ppm Cr(VI) concentration was determined to be optimal for further reduction studies.

Recyclability studies were performed for Cr(VI) reduction studies to further assess the stability of the composite. After centrifugation, the CT/BV@Fe-MOF catalyst can be separated from the Cr(VI) suspension and thoroughly cleaned with deionized water. The recycle study revealed that the reduction efficiency was decreased by 7.5% after six recycles (Fig. 3e). To further support the stability of the catalyst, after recovery of the photocatalyst from the multiple cycles of reusability experiments, the water samples are subjected to ICP-MS analysis to authenticate the precipitation of photocatalyst material on experimental conditions. The analysis reveals that no significant amount of Bi, Co or Fe ions were found, which confirms that no dissolution or precipitation of composite were noticed in the system. This suggested that the prepared composite appears to have a sound reduction efficiency even after six cycles of repeated application, supporting its practical utility in real-time applications.

To further understand the main reactive oxygen species involved during the photocatalytic reduction of Cr(IV), various radical quenchers, such as isopropyl alcohol (IPA), triethanolamine (TEOA), and p-benzoquinone (BQ) have been used. The results revealed that the relative changes in the Cr(VI) reduction efficiency are 34.89%, 84.91%, and 59.91% for the IPA, TEOA, and BQ scavengers, respectively (Fig. 3f). It is noticed that in the presence of TEOA, the photocatalytic system containing Cr(VI) showed a slight decrease in the efficiency profile. However, IPA and BQ relatively inhibit the Cr(VI) reduction efficiency, denoting the involvement of h+ and ∙O2− as the main reactive species during reduction process.

3.3 TCL degradation kinetics, DFT, and pathway

The degradation kinetics of TCL by the as-prepared catalysts were carried out in a laboratory-made photoreactor equipped with a 300 W Xe lamp as a visible light source. Firstly, the factors that influences the degradation of TCL have been investigated, and various pH ranges were studied, ranging from 3.0 to 11.0. The results indicated that high degradation efficiency has been identified in acidic conditions that turn into faster kinetics (Fig. 4a). The higher degradation efficiency at pH 5 can be attributed to the strong attraction forces between the negatively charged TCL molecule and positively charged composite surface. In addition, at lower pH values, the positively charged composite can provide more active sites for TCL adsorption and generate more reactive oxygen species for further degradation. The lower efficiency at higher pH values can be due to the repulsive force between the negatively charged TCL molecule and the composite. This leads to lower degradation efficiencies than the acidic values [57, 58].

Fig. 4
figure 4

Influence of (a) solution pH and b photocatalyst quantity on TCL degradation; c-d kinetics of TCL degradation under optimized experimental conditions; e reusability study, and f radical trapping experiments

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The influence of catalyst loading on TCL degradation has been further investigated, and the results indicated a significant increase from 51.6% to 97. 5% when the catalyst loads are at 25 mg and 50 mg, respectively (Fig. 4b). Interestingly, further increase in catalyst loading tremendously reduced the degradation of TCL in which at 75 mg of catalyst the efficiency was about 73.1%, while at 100 mg it further reduced to 61.4%. This can be attributed to the fact that at the optimum catalyst loading (50 mg), the composite has higher active sites for TCL adsorption and provides higher ROS generation [59]. The lower efficiencies at higher catalyst loads could be attributed to the fact that after the optimum concentration, the excess amount of catalyst will accumulate in the solution and minimize the generation of ROS. In addition, the excess amount of catalyst could also inhibit the light access to the composite particles to generate ROS during visible light irradiation [60, 61].

Furthermore, all the prepared catalysts showed insignificant degradation efficiencies under dark conditions, confirming that the catalyst showed higher degradation under visible light. The results showed that at the optimum conditions, the CT/BV@Fe-MOF composite showed the highest TCL degradation up to 97.5% under visible light irradiation (Fig. 4c). It is worth mentioning that the pure CoTiO3, BiVO4, CoTiO3@Fe-MOF, and BiVO4@Fe-MOF have shown lower degradation efficiency because they have high band gap values compared to the composite for efficient charge carriers’ generation and separation during visible light irradiation. Thus, it is confirmed that at the optimum conditions (pH 5, catalyst load = 50 mg), the as-prepared CT/BV@Fe-MOF composite showed the highest TCL degradation which is attributed to the higher number of active sites and optimum energy band gap, that facilitates the higher amounts of ROS (OH• − and O2• −) leads to efficient TCL degradation.

Additionally, the photocatalytic degradation of TCL has been examined as a function of their concentration, with a first-order kinetics demonstration (Fig. 4d). Furthermore, the study showed that the dosages up to 30 ppm could be degraded within a significant timeframe under ideal conditions, but as the concentration is increased, a significant decrease in catalytic activity was observed. The degradation process was slow at higher TCL concentration due to the reduction in number of active sites on the photocatalyst.

Recycling experiments were carried out at the optimum conditions to further assess the efficacy and stability of the as-prepared CT/BV@Fe-MOF composite. The results showed that the catalyst could retain its activity up to 90% even after 6 consecutive recycles (Fig. 4e). This suggested no significant reduction in catalytic activity and structural changes in the prepared catalyst. Thus, it is further confirmed that the synthesized composite has good stability and activity for practical applications.

Quenching experiments were carried out in our study to investigate the main ROS responsible for the TCL degradation. The results showed that the TCL degradation efficiency was around 97.5% without adding any radical trapping agents. However, after the addition of IPA as OH• − trapping agent, the TCL degradation efficiency has drastically dropped to 26%. After the addition of BQ as O2• − trapping agent it was decreased to 41% (Fig. 4f). In comparison, with the addition of TEOA as h+ trapping agent, and NaN3 as 1O2 trapping agent, the TCL degradation efficiencies were around 84% and 80.8%, respectively. Thus, these results strongly suggested that OH• − and O2• − are the dominant ROS for efficient TCL degradation by the as-prepared CT/BV@Fe-MOF composite [62, 63].

The geometric properties of TCL have been optimized through DFT calculations conducted in the gaseous phase using the B3LYP/6-31G (d,p) method (Fig. 5a). It is evident from the results that the TCL molecule exhibits three-dimensional geometric orientations. HOMOs and LUMOs have been used to evaluate the locations of TCL that are affected by the electrophilic and nucleophilic species, respectively. As shown in Fig. 5b, the HOMO & LUMO orbitals are systematically arranged, revealing a bandgap energy of 4.19 eV. The HOMO orbital of TCL is primarily associated with the amine, hydroxyl, and amide groups. In contrast, the LUMO orbital is associated with the aromatic groups, indicating that these areas of TCL are vulnerable to ROS attack.

Fig. 5
figure 5

a Optimized structure of TCL, b HOMO–LUMO energy of TCL and c Electrostatic potential maps of TCL molecule; d Plausible photocatalytic degradation pathway of TCL by the CT/BV@Fe-MOF nanocomposite

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Additionally, Fig. 5c illustrates the TCL molecular electrostatic potential map (ESP) diagram. According to the electron density analysis, sites with a higher charge density are more vulnerable to ROS attacks. ESP diagram further shows a negatively charged region that is susceptible to electron deprivation, indicating that TCL atoms within this region are vulnerable to attack by negatively charged O2– molecules.

To further understand the detailed degradation pathway of TCL by the as-prepared CT/BV@Fe-MOF composite, the LC/ESI–MS analysis was carried out at positive ion mode, and the corresponding path is shown in Fig. 5d. From the results, it is evident that there are two possible degradation pathways. In pathway-I, firstly, the TCL molecule can be attacked by the O2• − and lose the methyl group to generate DP-1 (m/z = 414.14), followed by further attack of OH• − or O2• − for deamination to generate DP-2 (m/z = 401.15). The intermediate structure can be further attacked by O2• − to lose the N-methyl group to cause DP-3 (m/z = 373.12), followed by the loss of hydroxyl groups to generate DP-4 (m/z = 357.16), which further undergoes ring opening process by the attack of OH• − or O2• − to generate harmless smaller molecules, CO2 and H2O (Fig. 5d). In the pathway-II, the TCL could be attacked by the OH• − and undergoes hydroxylation to generate DP-9 (m/z = 460.15), followed by further attack with OH• − to generate DP-10 (m/z = 428.86) and DP-11 (m/z = 396.17). The O2• − could further attack the intermediate products for deamination to cause DP-12 (m/z = 353.11), followed by ring opening reactions to generate smaller degradation products, which can finally convert into CO2 and H2O. To monitor the extent of TCL molecules mineralization, samples are subjected to TOC analysis. A TOC plot reveals that the total organic content of the drug samples decreases exponentially with increasing irradiation time. As a result of irradiating the CT/BV @Fe-MOF photocatalyst under optimal conditions, the residual TOC is observed to be 2.0 ppm within 1.5 h of photocatalytic irradiation (Fig. S4). As a result of this decrease, it can be concluded that the prepared nanocomposite is capable of effectively degrading the organic group in tetracycline and mineralizing it into CO2 and H2O.

3.4 Toxicity assessment

Toxicity assessment analysis is essential to understand the efficacy of the degradation process in detoxifying the TCL molecule during the degradation process by the as-prepared CT/BV@Fe-MOF composite. The toxicity analysis of TCL and its degradation products was assessed by the Toxicity Estimation Software Tool (TEST) [64, 65]. From the results, it showed that the TCL LC50 for fathead minnows suggesting it very toxic, and majority of the degradation products showed less toxicity compared to pure TCL (Fig. 6a). In the similar manner the LC50 of daphnia magna for TCL was found to be around 5.6 mg/L and with increase in degradation process the generated products showed decreased toxicity towards daphnia magna (Fig. 6b). Furthermore, the bioaccumulation factor for most of the developed degradation products has been reduced significantly compare to TCL, while DP-6 and DP-7 showed highest bioaccumulation with values BCF = 10.8 and BCF = 11.3, respectively (Fig. 6c). The degradation products DP-7, DP-11, DP-12, DP-13, and DP-14 showed negative mutagenicity while all other degradation products showed positive (Fig. 6d). Moreover, the developmental toxicity results revealed that except DP-2, DP-10, DP-11, and DP-12 all other degradation products showed lower developmental toxicity than pure TCL molecule (Fig. 6e). In general, it is worth to mention that although the toxicity of the TCL molecule has been significantly reduced during degradation process by the CT/BV@Fe-MOF composite, majority of the degradation products still hold their toxic nature.

Fig. 6
figure 6

a Acute toxicity on Fathead minnow and b Daphnia magna; c bioaccumulation factors, d mutagenicity and e developmental toxicity of intermediates generated upon TCL degradation by the CT/BV@Fe-MOF composite system

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3.5 Photocatalysis mechanism

To determine the band edge potentials of the valence and conduction bands of CoTiO3, BiVO4, and MIL-53 (Fe), Mulliken electronegativity empirical Eqs. (2) and (3) were applied.

$${\mathrm E}_{\mathrm{VB}}=\chi-{\mathrm E}_{\mathrm e}+0.5{\mathrm E}_{\mathrm g}$$
(2)
$${\mathrm E}_{\mathrm{CB}}={\mathrm E}_{\mathrm{VB}}-{\mathrm E}_{\mathrm g}$$
(3)

Here, EVB and ECB represent the band edge potentials of the valence and conduction bands, respectively, while χ is the electronegativity of the semiconductors, CoTiO3 (5.76), BiVO4 (6.04), and MIL-53 (Fe) (5.55 eV). The Ee represents the energy of free electrons on the hydrogen scale (Ee = 4.5 eV), and Eg is the band gap energy determined by UV–Vis-DRS analysis for BiVO4, CoTiO3, and MIL-53(Fe), which are 2.65, 2.50, and 2.30 eV, respectively. Accordingly, the ECB values for CoTiO3, BiVO4, and MIL-53 (Fe) were observed to be 0.01, 0.22, and -0.1 eV, respectively. Following this, the obtained EVB values are calculated as 2.51, 2.87, and 2.2 eV, respectively. All of these values are in consistent with the existing literature.

The results mentioned above demonstrated the formation of a unique heterojunction between CoTiO3, BiVO4, and MIL-53 (Fe) semiconductors, illustrating a traditional heterojunction-type photocatalytic mechanism (Fig. 7). In the presence of visible light irradiation, both MIL-53 (Fe) and CoTiO3 have the potential to be excited since their band gaps lie within the visible light range. The photoexcited electrons from the conduction band (CB) of MIL-53 (Fe) can easily migrate to the conduction band (CB) of CoTiO3, which further transfers to the conduction band (CB) of BiVO4. In addition, the photogenerated holes move from the valence band (VB) of BiVO4 to the valence band (VB) of CoTiO3. Thus, an efficient separation of the photogenerated charge carriers could be possible to generate a more significant number of ROS such as OH• − and O2• −. In the meantime, the separated e- could efficiently react with the O2 to produce O2•—on its surface, followed by the generation of h+ at the VB of CoTiO3 that could efficiently be further utilized for photochemical reactions to create OH• -. As a result, electron/hole recombination is reduced to a greater extent, and reactive oxygen species are formed, encouraging the photocatalytic degradation of Tetracycline and reduction Cr(VI). As summarized in Table 2, the proposed photocatalyst has exhibited superior performance compared to other reported materials in the literature for Cr(VI) reduction and TCL mineralization, respectively.

Fig. 7
figure 7

Photocatalytic degradation mechanism of CT/BV@Fe-MOF nanocomposite under visible light

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Table 2 Comparison of literature reports on the photocatalytic reduction and degradation of Cr(VI) and TCL drug with the proposed method
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4 Conclusion

In summary, the present study reports the preparation of a novel CoTiO3/BiVO4@MIL-53(Fe) nanocomposite that can be used as an effective visible light photocatalyst for the simultaneous reduction of Cr(VI) and degradation of Tetracycline. Different techniques have been used to characterized the photocatalyst's structural morphology, crystallinity, surface area, and porosity. The heterojunction construction promotes both the separation of photogenerated carriers and the retention of photogenerated carriers with a high oxidation/reduction capability. Photocatalytic performances of CT/BV@Fe-MOF is superior to those of pristine CoTiO3, BiVO4, and MIL-53 (Fe). Through trapping experiments, OH• and O2• − have been proven to be the main active substances to attack contaminants synergistically. This work reveals a new approach for developing high-performance visible light photocatalysts with enhanced active sites, larger surface areas, improved photogenerated carrier separation efficiency, and a high oxidation–reduction capability during inorganic and organic pollutants degradation process.