A Novel TiO₂-Cuttlebone Photocatalyst for Highly Efficient Catalytic Degradation of Tetracycline Hydrochloride

Abstract

The harmful effects of antibiotics on aquatic environments have become a growing concern of modern society. Developing high-performance photocatalysts capable of degrading antibiotics under solar light is, therefore, crucial. In this study, TiO₂-cuttlebone composites are prepared via the sol–gel method, to produce carbonate radicals (•CO₃⁻) under solar light irradiation. The •CO₃⁻ radicals exhibit high selectivity for the degradation of tetracycline hydrochloride (TC). Compared to TiO₂ alone, the TiO₂-cuttlebone composite demonstrates excellent solar-driven photocatalytic activity for TC degradation in both freshwater and seawater. The reaction pathways of TC degradation in seawater are elucidated using HPLC-MS/MS analysis. Moreover, a TiO₂-cuttlebone self-suspending photocatalyst device is fabricated using 3D printing technology and low-temperature deposition methods, with aluminum–plastic (AP) as a substrate. This innovative device is easily recyclable from photocatalytic solutions while maintaining high stability, making it highly desirable for practical applications.

1. Introduction

Over the past few decades, the rapid development of the aquaculture industry has led to the accumulation of significant amounts of antibiotics in seawater [1,2,3,4]. Tetracycline hydrochloride (TC), a common antimicrobial agent used in aquaculture, presents significant environmental hazards due to its low biodegradability and high stability [5,6,7]. The concentration of TC in aquaculture water, including freshwater and seawater, is generally in the range of μg/L⁻¹ to mg/L⁻¹. Although these levels are relatively low, the entry of TC into the food chain through ecological cycles results in bioaccumulation, posing great risks to the environment and to human health [8]. Therefore, the development of efficient and sustainable technologies to remove TC from both freshwater and seawater is urgently needed.

The application of photocatalysis for environmental remediation has attracted increasing research attention over the past decades [9,10,11,12]. Among these efforts, the photocatalytic degradation of TC has been demonstrated over a series of TiO₂ [13,14], WO₃ [15,16], and C₃N₄ [17,18] nanocomposites. Nevertheless, their photocatalytic efficiency is commonly limited by the complex molecular structure of TC. In a previous work, we reported a CaCO₃/TiO₂ composite photocatalyst that demonstrates excellent activity regarding the degradation of TC [19]. This high performance is attributable to the reaction between photocatalytically generated •OH and holes with CO₃²⁻, which produces carbonate radicals (•CO₃⁻) that exhibit unique activity and selectivity toward TC degradation. Cuttlefish bone is an abundant, low-cost, environmentally friendly, and non-toxic material in the ocean that is particularly rich in CaCO₃. Therefore, combining TiO₂ with cuttlebone is a promising approach for generating a good deal of •CO₃⁻ free radicals under light irradiation, thereby accelerating TC degradation. However, no reports currently exist in the open literature on this topic.

Moreover, most catalysts developed for photocatalytic degradation are powder-based, which makes them difficult to recover in practical aquaculture applications and may potentially cause secondary pollution. Anchoring nanometer photocatalysts onto bulk substrates is considered to be an effective strategy to address these challenges [20,21,22,23,24,25]. In this regard, we have designed and fabricated self-suspending aluminum–plastic (AP)/semiconductor photocatalyst devices for the degradation of TC [26]. Compared to powder systems, the device maximizes light utilization, quickly replenishes oxygen during the photocatalytic process, and facilitates the generation of active •OH and •O₂⁻ species. Furthermore, the conductive aluminum accelerates electron separation and reduces the recombination of charge carriers [26,27]. In this context, the fabrication of TiO₂-cuttlebone composite and further immobilization of the designated catalysts on the AP substrate are anticipated to enhance both photocatalytic activity and cyclic stability.

Bearing these considerations in mind, for the present study, we first synthesized TiO₂-cuttlebone (TCb) composites via the sol–gel method for the degradation of TC (5 and 30 mg/L) in freshwater and seawater under solar light and natural light irradiation, respectively. It is clear that the combination of cuttlebone with TiO₂ remarkably enhances TC degradation efficiency compared to pure TiO₂, primarily due to the generation of a large number of •CO₃⁻ radicals under light irradiation. A self-suspending photocatalyst device was then prepared using an AP substrate, through 3D printing technology and low-temperature deposition. TCb nanoparticles were immobilized on the surface of the AP plate. This floating device improves light utilization and the oxygenation rate of the photocatalyst. The catalytic activity of the self-suspending AP-TCb device is notably higher than that of TCb powder, under the condition of equal TCb mass. This work provides new insights into the preparation of selective oxidation photocatalytic materials for TC degradation and the fabrication of photocatalyst devices, with promising applications in real-world seawater treatment.

2. Results and Discussion

Figure 1a shows the XRD partners of the prepared catalysts. TiO₂ exhibits diffraction peaks at 25.3°, 37.8°, 48.1°, 53.9°, and 55.1°, corresponding to the (101), (004), (200), (105), and (211) crystal planes of anatase TiO₂ (JCPDS No. 21-1272), respectively [28]. The XRD pattern for cuttlebone powder shows diffraction peaks at 26.3°, 27.3°, 33.2°, 36.3°, 38.0°, 38.5°, 41.3°, 43.0°, 45.9°, 48.5°, 50.3°, and 52.5°, which are assignable to orthorhombic CaCO₃ (JCPDS No. 05-0453) [29]. Additionally, the XRD pattern of the resulting TCb nanocomposite reveals diffraction peaks corresponding to both anatase TiO₂ and orthorhombic CaCO₃, indicating that the sol–gel method, followed by calcination, leads to the successful hybridization of TiO₂ with cuttlebone powder.

Figure 1b displays the FT-IR spectra of the cuttlebone, TiO₂, and TCb samples. In the TiO₂ spectrum, a peak at 3422 cm⁻¹ corresponding to the stretching and bending vibrations of the O-H bond is observed. In addition, a wide band in the range of 400–800 cm⁻¹ is attributable to the typical vibration of the O–Ti–O bond within TiO₂. A peak at 1383 cm⁻¹ can be assigned to the –OCH₂CH₃ groups, which might result from the incomplete hydrolysis of titanium tetraisopropoxide during the synthetic processing of TiO₂ sol [30]. Pure cuttlebone exhibits four characteristic peaks at 1475 cm⁻¹, 1083 cm⁻¹, 875 cm⁻¹, and 712 cm⁻¹, all of which can be attributed to the stretching vibrations of C–O bonds [31]. In the TCb composite, the main peaks from both TiO₂ and cuttlebone are apparent, demonstrating the coexistence of the two components.

The presence of Ti, O, Ca, and C elements in the TCb nanocomposite is confirmed by XPS analysis (Figure 1c). In Figure 1d, the high-resolution Ti 2p XPS spectrum shows Ti 2p_1/2 and Ti 2p_3/2 peaks at 465.0 and 459.3 eV, respectively. The O 1s XPS peak can be deconvoluted into three components at 532.8, 531.9, and 530.7 eV (Figure 1e), which are ascribable to surface-adsorbed oxygen species, carbonate-like species, and lattice oxygen, respectively [32]. In Figure 1f, the peaks at 350.9 and 347.3 eV are attributable to Ca 2p_1/2 and Ca 2p_3/2, confirming the +2 oxidation state of Ca in the composite [19]. The XPS spectrum in the C 1s region is deconvoluted into peaks at 289.7, 289.3, 286.3, and 284.8 eV, attributable to CO₃²⁻, O=C-O, C-O, and C-C bonds, respectively (Figure 1g). The existence of the C–C group is consistent with the incomplete hydrolysis of titanium tetraisoproxide during the synthesis process of the TiO₂ sol. This finding aligns with the results from the FT-IR analysis. Overall, these results provide solid evidence for the construction of the TiO₂-cuttlebone hybrid composite.

The micromorphology of TCb was examined using SEM. As shown in Figure 2a, TCb is mainly composed of clusters. To further investigate the detailed structural and morphological characteristics, transmission electron microscopy (TEM) analysis was performed (Figure 2b,c). The result is consistent with the SEM image. Moreover, the nanoparticles are distributed in clusters with a width and length of approximately 100 nm and 1 μm. HR-TEM analysis was further conducted on the TCb sample (Figure 2c–e). The measured lattice fringe distances of 0.35 and 0.19 nm correspond to the (101) and (200) crystal planes of anatase TiO₂, respectively. The size of the TiO₂ nanoparticles was measured to be about 7 nm. Furthermore, the elemental mapping analysis of the TCb sample revealed the presence of Ca, Ti, C, and O, as shown in Figure 2f,g. Ca and Ti elements were evenly distributed throughout the composite. The EDS result (Figure S1) shows that the content of Ti and Ca elements is 10.55% and 12.08%, which is close to the feed ratio of the synthetic process. These findings further confirm the immobilization of TiO₂ nanoparticles on the cuttlebone component.

The specific surface area and pore structure of the synthesized samples were evaluated via the N₂ adsorption–desorption technique, as shown in Figure 3a. Both the TiO₂ and TCb samples exhibit type IV curves, indicative of the typical adsorption behavior of mesoporous structures. Based on the Brunauer–Emmett–Teller (BET) isotherms, the specific surface areas of the cuttlebone, TiO₂, and TCb samples were calculated to be 8.0, 167.4, and 107.5 m²/g, respectively. Furthermore, the BJH patterns (Figure 3b) revealed that TiO₂ has an average pore size of 7.2 nm, with a narrow pore size distribution similar to that observed for the TCb sample.

Figure 3c displays the UV-vis DRS spectra of the prepared samples. Bare TiO₂ shows strong absorption in the region of 200–398 nm, which is consistent with its band gap of 3.1 eV [33]. In comparison, the UV-vis DRS spectrum of the TCb nanocomposite exhibits enhanced light absorption in the 400~800 nm region, which could be favorable for photocatalytic reactions.

The photocatalytic activities of the synthesized catalyst were initially investigated by the degradation of TC in freshwater, as shown in Figure 4. Before photocatalytic degradation, the catalyst was stored in the dark to adsorb TC for 30 min to establish absorption–desorption equilibrium. After 5 min of light irradiation, 96.4% of TC (5 mg/L) was degraded over the TCb catalyst (Figure 4a). In contrast, no degradation of TC was observed in the dark, and only 15.4% of TC was eliminated under light irradiation without a catalyst. Bare TiO₂ showed inferior photocatalytic performance, with 71.9% of TC degradation under the same conditions. The obviously higher photocatalytic activity of the TCb nanocomposite compared to bare TiO₂ cannot be ascribed to differences in the specific surface areas since the TCb nanocomposite has a smaller surface area (107.5 m²/g) than bare TiO₂ (167.4 m²/g). Instead, the enhanced photocatalytic activity of the TCb nanocomposite suggests that cuttlebone plays a key role in promoting TC degradation.

The ratio of TC to photocatalyst in the reaction system also influences performance. When the concentration of TC in the reaction system increased to 30 mg L⁻¹ while the amount of photocatalyst remained constant, 79.7% and 92.0% of the TC was degraded within a light irradiation of 30 min over TiO₂ and TCb, respectively (Figure 4b). This is less efficient compared to the degradation of 5 mg L⁻¹ TC. The TCb nanocomposite also exhibited high stability during the photocatalytic degradation of TC (30 mg L⁻¹). Recent progress in the photocatalytic degradation of TC by TiO₂-based materials in recent years is shown in Table S1 [19,29,34,35,36,37,38], indicating that TCB has a high degradation efficiency for TC. As presented in Figure 4c, the photocatalytic degradation of TC in freshwater was successfully repeated five times without obvious changes in activity.

LC-QTOF-MS was employed to investigate the intermediates generated in the system, as shown in Figure 5a–c. A possible degradation pathway of TC is proposed in Scheme 1. Firstly, CO₃⁻ initiates an oxidative attack on the target molecule, generating the intermediate product of M₁ (m/z 431) [39]. Then, M₁ undergoes deamination, demethylation, and addition reactions to form M₂ (m/z 417) and M₃ (m/z 305) [40]. Subsequently, M₃ undergoes a series of ring-opening reactions, gradually decomposing into smaller molecules, including M₄ and M₅. These smaller molecules eventually decompose further, yielding H₂O, CO₂, and other inorganic substances. Concurrently, CO₃⁻ selectively oxidizes the target molecule, producing another intermediate product of M₆ (m/z 459), as described in previous research [41]. M₆ undergoes demethylation, deamination, deprotonation, and addition reactions, and a six-membered ring is opened to generate M₇ (m/z 411) and M₈ (m/z 344). M₇ and M₈ continue to undergo ring opening and demethylation reactions, further decomposing into smaller molecular compounds. Finally, these smaller molecules undergo a series of oxidative degradations, ultimately decomposing the target molecule and completing the degradation process.

To explore the effect of the photocatalytic mechanism of TC over TCb, BQ, AO, IPA, and 4-cp were employed as scavengers of •O₂⁻, h⁺, •OH, and •CO₃⁻ and added into a TC aqueous solution. As shown in Figure 6a, •O₂⁻ and h⁺ play an important role in the degradation of TC over TiO₂ instead of •OH and •CO₃⁻ radicals. This result is consistent with our previous research [29].

Compared with TiO₂, the TCb composite shows a different mechanism of TC degradation (Figure 6b). An obvious decrease in TC degradation efficiency occurs after the AO scavenger is added to the system. When the BQ scavenger is added, there is a certain effect on degradation. However, the negative effect of BQ and AO on TC degradation is weaker than that of TiO₂. This may be due to the consumption of a portion of the species of •O₂⁻ or h⁺ to facilitate the conversion of CO₃²⁻ into •CO₃⁻. In the case of IPA, there are no obvious effects on the catalytic activity of TCb. Furthermore, TC degradation efficiency is remarkably reduced when 4-CP is added, indicating that the •CO₃⁻ radical is significant in this reaction [19]. Furthermore, ESR measurements (Figure S2) demonstrate the existence of the •CO₃⁻ radical in the photochemical system of TCb.

The above observations clearly indicate that the degradation of TC is a photocatalytic process over TiO₂, with cuttlebone acting as a promoter to enhance photocatalytic efficiency. Under solar light irradiation, TiO₂ is excited and produces electrons (e⁻) and holes (h⁺). The e⁻ can activate O₂ to generate •O₂⁻ species, while the h⁺ contribute to the formation of •OH radicals. The generated •OH, •O₂⁻, or h⁺ species can then react with CO₃²⁻/HCO₃⁻ to produce selective •CO₃⁻ radicals [29]. This leads to the synergistic degradation of TC over the TCb hybrid composite, resulting in improved photocatalytic performance.

Based on the excellent photocatalytic performance regarding TC degradation in freshwater, the photocatalytic activity for the degradation of TC (5 mg/L) in seawater over TiO₂ and TCb has also been investigated, due to its great significance for environmental protection. As displayed in Figure 7a, the TC degradation efficiency reaches 88.5% and 97.8% within 20 min with TiO₂ and TCb, respectively. Moreover, the photocatalytic degradation of TC in seawater can be cycled five times without a noticeable loss in activity (Figure 7b). These results highlight the great potential of TCb for the elimination of TC in seawater.

To promote the practical application of the prepared TCb, we further immobilized the catalyst on an AP substrate to create a self-suspending AP-TCb photocatalyst device (for more details, please refer to the experimental section). Figure S3 shows the SEM image of the cross-section of AP-TCb. As depicted, TCb is distributed over the surface of AP, with a thickness of about 3 μm.

The photocatalytic performance of AP-TCb has been evaluated for TC degradation in freshwater and seawater under natural light, and an indicative scheme of experiment arrangement is proposed (Scheme 2). As illustrated in Figure 8a,b, after light irradiation for 20 min, the degradation efficiency of TC was 98.34% in freshwater and 96.4% in seawater. Compared to TCb powder, the AP-TCb device exhibited significant enhancements for TC degradation in both freshwater and seawater, which can be attributed to the following three major factors: (a) increased light-capture efficiency, leading to greater light utilization; (b) the timely replenishment of oxygen consumed in the photocatalytic process; and (c) the presence of Al in the AP, which accelerates electron migration, promoting the separation of electrons and holes [26]. These results demonstrate the great potential of this AP-TCb self-suspending device for practical water treatment in both freshwater and seawater environments.

3. Methods and Materials

3.1. Chemical and Reagents

Tetracycline hydrochloride (96%), benzoquinone (BQ, 99%), ammonium oxalate (AO, 99.8%), isopropyl alcohol (IPA, ≥99.5%), 4-chlorophenol (4-CP, 99%), and 5,5-dimethyl-1-pyrroline-N-oxide (DMPO) were purchased from Aladdin (Shanghai, China). Titanium tetraisopropoxide (C₁₂H₂₈O₄Ti, 97%) was obtained from Sigma-Aldrich (St. Louis, MO, USA). Waste cuttlebone was obtained from a seafood market in Fuzhou, China. Seawater was obtained from Pingtan, China.

3.2. Synthesis Process

The raw cuttlebone powder was prepared via a grinding method, followed by sieving through a 100-mesh screen. TiO₂-cuttlebone nanocomposites (TCb) were synthesized via the sol–gel method, followed by a calcining process. In brief, titanium tetraisopropoxide was first hydrolyzed under acidic conditions. The resulting suspension was dialyzed to a pH of ~4. According to our previous work, the calcination of 100 mL of titanium glue can obtain 3 g TiO₂. Then, the raw cuttlebone powder (1.8 g) was added to the obtained transparent TiO₂ sol (40 mL), dehydrated in a microwave oven, and calcinated at 300 °C for 5 h. The resulting product was washed with ethanol and dried overnight at 60 °C. For comparison, bare TiO₂ and cuttlebone were synthesized using the same procedure without adding raw cuttlebone or TiO₂ sol.

The self-suspending AP substrate was fabricated by 3D printing, according to our previous work [26]. For TCb immobilization, the AP substrate was washed with ethanol several times and dried. Its initial mass was recorded as m₁. TCb (0.10 g) was dispersed in 50 mL of anhydrous ethanol and ultrasonicated for 30 min to obtain a TCb suspension with a concentration of 2 g/L. The suspension was then poured into a Petri dish containing the AP substrate and dried in an air-circulating oven at 130 °C for 5 h. Finally, the impregnated AP substrate was ultrasonically cleaned in deionized water for 60 min to remove any loose TCb from the AP surface, followed by drying at 60 °C to obtain the AP-TCb device. The final mass was recorded as m₂. The amount of TCb loaded onto the AP substrate was determined by calculating the difference between m₂ and m₁.

3.3. Characterization and Analytical Methods

X-ray diffraction (XRD) patterns were studied with a Bruker D8 Advance X-ray diffractometer using Cu K_α radiation (40 kV, 40 mA, λ = 1.5418 Å) in a 2θ range of 5–90° with a 0.02° step size. X-ray photoelectron spectroscopy (XPS) was conducted with a PHI Quantum 2000 system, using a monochromatic Al K_α source. The specific surface areas were obtained at −196 °C via a BELSORP-mini II nitrogen adsorption–desorption apparatus. The Fourier transform infrared (FT-IR) spectra were determined in transmittance mode with a resolution of 4 cm⁻¹, using a Nicolet Is 10 FT-IR spectrometer. The surface morphology of the samples was recorded with scanning electron microscopy (SEM, Regulus 8100, Tokyo, Japan) and transmission electron microscopy (TEM) (FEI, Talos F200X, Boston, MA, USA). UV-vis diffuse reflectance spectroscopy (DRS) measurements were determined using a Varian Cary 500 spectrometer equipped with an integrating sphere, employing BaSO₄ as a reference. A Bruker ESP 300E spectrometer was employed to measure the electron spin resonance (ESR) signals of the radicals that were spin-trapped by DMPO.

3.4. Photochemical Measurements

The photocatalytic reactions were performed under simulated solar light irradiation. The irradiance was determined to be 269 mW·cm⁻², using a PL-MW 2000 photoradiometer (Beijing Perfect Light Co.; Beijing, China) at a distance of 10 cm from the 300 W Xenon lamp to liquid level and an optical power of 0.845 W, equipped with an AM 1.5 G simulated solar filter (300 W xenon lamp, PLS-SXE300, Beijing Perfect Light Co.; Beijing, China) and optical equalizer (PLS-LA320A, Beijing Perfect Light Co.; Beijing, China). Typically, a glass vessel with a water-cooling system (25 °C) was used to mix 50 mg of photocatalyst and 100 mL of TC solution at a specific concentration in the dark for 30 min, to achieve absorption-desorption equilibrium. Subsequently, the suspension was exposed to Xe lamp irradiation within the wavelength range of 200–1200 nm. A 4 mL sample from the reactor was filtered through a 0.45 μm filter membrane for analysis, using either an Agilent high-performance liquid chromatography spectrometer (1260 infinity LC) (Hong Kong, China) or a Shimadzu UV-1750 UV-vis spectrophotometer (Kyoto, Japan). The wavelength for TC detection was set at 357 nm.

The AP model was designed to be a disc-shaped structure with a radius of r₁ = 35 mm, a height of h₁ = 1.2 mm, a wall thickness of t₁ = 0.4 mm, and an infill ratio of η = 58.71% for the central infill layer. The model featured 102 circular holes with a diameter of 1 mm each to facilitate water flow. Using a densitometer, the density of the prepared AP filament was measured to be ρ₀ = 1.155 g·cm⁻³. If the model was filled to 100% infill, the device would fail to achieve self-suspension, due to its density being greater than that of water (1 g·cm⁻³). Therefore, the central part of the model needed to be filled at a specific infill ratio to bring the overall density of the device close to that of water. We set the density of the model to be ρ₁ ≈ 1 g·cm⁻³, calculated using the following equation:

$${\rho }_1={\displaystyle \frac{m_1+m_2}{V_0}}$$

(1)

where ρ₁ is the density of the model, m₁ is the quality of the model wall, m₂ is the quality of the middle fill layer, and V₀ is the volume of the model.

According to the mass formula, m = ρV, the volume formulae V = hπr², m₁ = 3.522 g, V₀ = 4.520 cm³, and the volume of the central infill layer, V₂ = 1.472 cm³. Therefore, using Equation (1), the mass of the central infill layer m₂ can be calculated to be 0.998 g. In the following equation:

$$\mathsf{\eta }={\displaystyle \frac{m_2}{{\rho }_0{V}_2}}\times 100\%$$

(2)

η is the filling rate of the middle filling layer.

According to our calculations, η = 58.71%.

4. Conclusions

In summary, we have successfully prepared a novel cuttlebone-TiO₂ (TCb) hybrid composite, as presented in the current work. Compared to pure TiO₂, the obtained TCb composite exhibits superior TC degradation efficiency under solar light and natural light irradiation in freshwater and seawater. The enhanced degradation rate is attributable to the production of highly selective •CO₃⁻ radicals in the reaction system. The degradation pathway of TC was explored by LC-MS analysis. To facilitate the recovery of the TCb photocatalyst in practical applications, an AP material was utilized as a substrate to prepare an AP-TCb self-suspending device. The device not only realizes the easy separation and recovery of the catalyst but also demonstrates obviously improved photocatalytic activity compared to the powder sample. This study offers a promising and convenient approach for treating low concentrations of antibiotics in real-world aquaculture wastewater.