First Utilization of Magnetically-Assisted Photocatalytic Iron Oxide-TiO₂ Nanocomposites for the Degradation of the Problematic Antibiotic Ciprofloxacin in an Aqueous Environment

Abstract

The emergence of antimicrobial resistance due to antibiotics in the environment presents significant public health, economic, and societal risks. This study addresses the need for effective strategies to reduce antibiotic residues, focusing on ciprofloxacin degradation. Magnetic iron oxide nanoparticles (IO NPs), approximately 13 nm in size, were synthesized and functionalized with branched polyethyleneimine (bPEI) to obtain a positive charge. These IO-bPEI NPs were combined with negatively charged titanium dioxide NPs (TiO₂@CA) to form magnetically photocatalytic IO-TiO₂ nanocomposites. Characterization techniques, including X-ray diffraction (XRD), transmission electron microscopy (TEM), scanning electron microscopy (SEM), infrared spectroscopy (IR), X-ray photoelectron spectroscopy (XPS), thermogravimetric analysis (TGA), electrokinetic measurements, and a vibrating sample magnetometer (VSM), confirmed the successful formation and properties of the nanocomposites. The nanocomposites exhibited a high specific surface area, reduced mobility of photogenerated charge carriers, and enhanced photocatalytic properties. Testing the photocatalytic potential of IO-TiO₂ with ciprofloxacin in water under UV-B light achieved up to 70% degradation in 150 min, with a degradation rate of 0.0063 min⁻¹. The nanocomposite was magnetically removed after photocatalysis and successfully regenerated for reuse. These findings highlight the potential of IO-TiO₂ nanocomposites for reducing ciprofloxacin levels in wastewater, helping curb antibiotic resistance.

1. Introduction

Antibiotics are among the crucial drugs essential for improving human health. However, the rise of bacterial resistance to antibiotics and their rapid adaptation have become a global issue [1]. Despite a record low in antibiotic usage recorded in 2020, it remains one of the top three health threats, according to the European Health Emergency Preparedness and Response Authority (HERA). Infections caused by antibiotic-resistant bacteria can lead to additional healthcare and productivity costs, potentially amounting to as much as 1.5 billion euros [2]. Between 2017 and 2021, infections caused by Klebsiella pneumoniae, a rod-shaped bacterium commonly found in the intestines that can infect the lungs, increased. These infections are often treated with ciprofloxacin [3].

Ciprofloxacin, a synthetic broad-spectrum antibiotic, is utilized for its bactericidal and bacteriostatic effects against a wide range of Gram-positive and Gram-negative bacteria [4,5]. This has resulted in its increased use in clinical treatments. However, due to its structural stability, ciprofloxacin cannot be completely metabolized in the human body. Studies have reported that about 40–60% of ciprofloxacin is excreted by humans in unmetabolized form [6]. Residues from both metabolized, unused, or expired ciprofloxacin-based drugs are released into the environment. For example, a study found that hospital wastewater in France contained ciprofloxacin concentrations ranging from 0.97 to 3.39 μg/L [7]. These residues bind to minerals and organic compounds, which accumulate in soil or disperse rapidly in water [8], leading to ecotoxicological effects [5]. Ciprofloxacin ingestion through water has been associated with unhealthy symptoms such as diarrhea, headaches, vomiting, and tremors [9,10]. Urgent action is required to remove the antibiotic from wastewater to safeguard underground water supplies.

Various investigations have focused on this issue, considering the physicochemical properties of water. Common procedures employed to eliminate antibiotics include coagulation [11], flocculation [12,13], sedimentation [14], filtration [15], and disinfection [16]. However, due to the formation of stable wastewater mixtures, those classical methods proved to be either ineffective or insufficient [17], having several shortcomings. Additional investigations using catalytic oxidation [18], electro-Fenton photocatalytic oxidation [19], and adsorption proved to be more efficient, especially adsorption using nanotechnology in combination with photocatalysis [17,20]. Photodegradation offers numerous appealing advantages over traditional methods, including being environmentally friendly, cost-effective, efficient, reusable, and scalable [21,22]. In comparison, for example, to the Fenton oxidation process, where large volumes of iron slurry are obtained without the possibility of reusability, photodegradation is the preferred method [23]. Photocatalysis utilizes sunlight or artificial light, and photocatalysts degrade persistent pollutants under light irradiation and contribute to a comprehensive and sustainable wastewater treatment solution. Advanced oxidation processes (AOPs) are particularly important and attractive for the degradation of pollutants in contaminated water. These processes generate strongly oxidizing radicals under UV light, especially hydroxyl radicals (OH·). These radicals are able to break down complex compounds into simpler, non-toxic substances, making AOPs an effective method for water purification [24].

Commonly used materials for photodegradation include metal oxides [25] such as titanium dioxide (TiO₂) [26], zinc oxide (ZnO), graphene oxide [27], iron oxide [28], and metal-organic frameworks [9]. Due to its cost-effectiveness and high efficiency, TiO₂ is the most commonly used material [29]. TiO₂, a semiconductor photocatalyst, is often used in heterogeneous processes to degrade pollutants in aqueous environments. TiO₂ NPs have remarkable properties, including a high surface-to-volume ratio, strong oxidizing power, and low toxicity. As a result, TiO₂ NPs are gaining ground in the field of photocatalysis and are being extensively studied for various photocatalytic applications. Importantly, a deeper understanding of TiO₂ photocatalysis has facilitated the development of new photocatalytic materials with improved functionalities, such as integration with magnetic materials [24,30]. However, the effectiveness of TiO₂ is hindered by its wide bandgap, which limits its efficiency in visible light and has reduced commercial interest in recent years. In fact, the controlled application and removal of TiO₂ nanocomposites is essential. A major challenge is the difficulty in removing the photocatalyst from the solution medium after use. Although various technologies, primarily filtration, have been used to remove TiO₂ after photocatalysis, they often result in loss or incomplete removal of TiO₂ [31,32]. Consequently, further research has focused on the immobilization of TiO₂ on supports such as glass plates [33], waste polystyrene [34], or magnetic nanomaterials [31,32,35,36,37,38,39,40] to address these challenges. In the field of TiO₂-based photocatalysts, considerable efforts have been made to incorporate magnetic nanomaterials into TiO₂-based nanocomposite structures. This integration aims to enable the rapid and efficient removal of photocatalytic materials from aqueous systems using an external magnetic field. The multifaceted approach here in this study emphasizes the potential of magnetic nanocomposites as advanced, viable, and versatile tools in addressing the complex challenges associated with wastewater remediation [41].

In particular, combining TiO₂ with other metal iron oxides, such as hematite (α-Fe₂O₃) [29,37,42,43,44], maghemite (γ-Fe₂O₃) [45], or magnetite (Fe₃O₄) [24,31,32,36,38,40,46,47,48], can improve its performance. Correspondingly, maghemite iron oxide is non-toxic and cost-effective, making it a viable option for scaling up wastewater treatment technologies [45,49,50]. The combination of TiO₂ and Fe₂O₃ has been found to be highly effective in the photo-oxidation of water, generating photo-induced holes that can last from milliseconds to seconds [44].

The synthesis/combination of iron oxides and TiO₂ nanocomposites typically involves multi-step processes that require high temperatures [36,51] or the use of toxic materials [26,52], posing challenges in terms of cost and environmental impact. Researchers are exploring alternative methods using charge complexation, sonication for activation, or incorporation of silicium dioxide [53] as a median layer to prepare TiO₂-iron oxide nanoparticles. Furthermore, under mild conditions with non-toxic solvents or activators, these nanoparticles were made using poly(N-isopropylacrylamide-co-methacrylic acid) and silver [54]. These advances aim to improve the feasibility and safety of nanocomposites in wastewater remediation. In the development of nanocomposite materials from two different NPs, hetero-agglomeration is an excellent choice due to its simplicity, lack of specific additional chemicals, and environmental friendliness. It is based on attractive forces between oppositely charged nanoparticles, resulting in composites that are irreversibly bonded together and exhibit high stability [40].

Most photocatalytic materials based on magnetic materials such as Fe₃O₄ or α-Fe₂O₃, or their combination with TiO₂, have primarily been utilized for degrading dyes [24,31,32,37,38,39,40,43,44,46,48,54,55,56,57] or other organic pollutants [35,36,51]. However, few studies have focused on the degradation of the antibiotic ciprofloxacin, followed by magnetic-assisted removal, as indicated by the cited research [33,34,47,58,59,60,61,62]. In this context, photocatalytic materials developed for the degradation of ciprofloxacin have mainly focused on materials such as TiO₂ [61], chlorophyll-sensitized and salicylic acid-functionalized TiO₂ NPs [58], Fe₂O₃ nanoparticles loaded on graphitic carbon nitride [59], waste polystyrene and TiO₂ composites [34], defective TiO₂-x nanomaterials [60], and biosynthesized silica-supported silver nanoparticles [62]. In addition, other materials, such as ZnIn₂S₄/CoFe₂O₄ p-n junction decorated biochar [63], Gd₂WO₆-loaded ZnO/bentonite nanocomposite [64], Fe₃O₄/CdS/g-C₃N₄ [65], CuFe₂O₄@methylcellulose-based magnetic nanobiocomposite [66], nano-TiO₂ immobilized polyvinylidene fluoride-based sponge beads [67] and others [68,69,70] have been used for photocatalytic degradation of ciprofloxacin, but no study uses the IO-TiO₂ nanocomposite for the degradation of ciprofloxacin, which highlights the unique contribution of the present work to the scientific community. In fact, the combined approach of photocatalytic degradation of ciprofloxacin and magnetic removal of the photocatalytic material after the process is relatively unexplored in the literature and requires further investigation. Currently, there are no studies addressing the hetero-agglomeration of positively modified IO-bPEI NPs with negatively charged TiO₂@CA for the photocatalytic degradation of ciprofloxacin in an aqueous environment, followed by the removal of the nanocomposite with an applied magnetic field. This gap in research clearly emphasizes the novelty and significance of the present study. Additionally, the innovative use of bPEI and citric acid in this combination has not been previously reported in the literature, making this approach both novel and more environmentally friendly. This method ensures that only the precise amount of TiO₂ is adsorbed onto the IO-bPEI, thus controlling the TiO₂ usage.

In this study, a precisely engineered IO-TiO₂ nanocomposite was developed by combining positively charged maghemite iron oxide NPs (i.e., IO-bPEI) with negatively charged TiO₂@CA NPs through a simple hetero-agglomeration method. This nanocomposite was applied for the photocatalytic degradation of ciprofloxacin under UV-B irradiation. After completion of the process, the nanocomposites could be easily removed using a permanent magnet. This represents a significant advance over existing state-of-the-art photocatalysis processes for the degradation of ciprofloxacin. Although this work was studied under UV irradiation, it lays the foundation for future studies in the visible light range. The carefully prepared nanomaterials were thoroughly investigated for their structure, morphology, composition, electrophoretic behavior, and magnetic properties using various methods such as ATR-IR spectroscopy, potentiometric titrations, XRD, XPS, SEM with EDXS, TEM, TGA, VSM, and BET, among others. Several factors affecting the degradation of ciprofloxacin under UV-B irradiation were investigated, and optimal conditions for efficient photodegradation were defined. The photocatalytic potential of the nanocomposite was also explored (i.e., UV-Vis reflectance spectra (UV-Vis-DR), photoluminescence (PL) emission, band gap). In addition, the recyclability and reusability of photocatalysts are crucial for the environmentally friendly and sustainable development of photocatalytic technology, although this is often overlooked. Therefore, the stability and reusability of nanocomposite were explored. Finally, plausible degradation mechanisms for ciprofloxacin using the developed nanocomposite are exhaustively discussed.

2. Materials and Methods

2.1. Materials

All the chemicals were used as received, without any further purification. The FeSO₄ × 7H₂O, Fe₂(SO₄)₃ × 7H₂O, and HCl (≥37%), branched PEI with an average MW of 25,000 Da, and ciprofloxacin with a purity of ≥98% (HPLC) were received from Sigma-Aldrich. The NH₄OH (25% aqueous solution) and NaOH (>98%) were purchased from Honeywell. The absolute EtOH (anhydrous) was received from CarloErba.

In this study, we used a commercially available nano-sized colloidal dispersion of TiO₂@CA, namely CCA100BS—a dispersion with neutral pH 7 (~25% TiO₂@CA according to the manufacturer). These NPs were industrially produced and supplied by Cinkarna, Metallurgical and Chemical Industry Celje, Inc. in Slovenia (Celje, Slovenia). The TiO₂@CA NPs in this stable dispersion were synthesized in a sol-gel process using metatitanic acid, a by-product of the sulfate synthesis process. The complete synthesis process is described in patent SI 23,501 A of Cinkarna Celje. Details can be found in the Supplementary Information. The magnet used for the magnetic separation was a nickel-plated NdFeB block magnet (size 46 × 30 × 10 mm) with an adhesive force of approx. 284 N (approx. 29 kg) and a magnetization of N40, which was purchased from supermagnete.

Ultra-pure water with a resistivity > 18 MΩ cm (Millipore, Burlington, MA, USA) was used for the preparation of all the aqueous systems in this work.

2.2. Synthesis and Functionalization of Magnetic Iron Oxide Nanoparticles with Branched Polyethylene Imine (IO-bPEI)

Magnetic iron oxide nanoparticles (IO NPs) were synthesized using a slightly modified method based on [71]. Iron sulfate salts were dissolved in 500 mL of MilliQ water under ambient air conditions at a molar ratio of Fe²⁺ to Fe³⁺ of 2.4:1. A diluted ammonium solution was then added to the mixture to adjust the pH to 3. This value was maintained for 30 min by continuous addition of the ammonium solution, resulting in the precipitation of iron hydroxides. Subsequently, 250 mL of 25% NH₄OH was added to form iron oxide MNPs, and the solution was magnetically stirred for another 30 min. To purify the synthesized MNPs, MilliQ water was prepared with diluted NH₄OH at pH 10. The MNPs were washed three times with 200 mL of this alkaline solution, once with pure MilliQ water with a separation time of about 1 min, and finally dispersed in MilliQ water.

Functionalization of bare IO NPs with branched polyethylenimine (bPEI) was performed under the assumption that 5 bPEI molecules would bind per 1 nm² of the MNPs’ surface. The required amount of polysaccharide was calculated based on a theoretical specific surface area of 95.4 m²/g for ~13 nm MNPs as reported by [71]. For the functionalization of 250 mg MNPs, 100 mg bPEI was required. To achieve this, a 0.4 wt% dispersion of MNPs (62.5 mL) was prepared, and the pH was adjusted to 8 with 0.1 M NaOH. The bPEI (0.2 wt%) was dissolved in MilliQ water, also adjusted to pH 8, and then slowly added to the MNP dispersion. At this pH, the amino groups of bPEI are protonated and positively charged, while the bare MNP carries a negative charge, which facilitates electrostatic interaction (as shown by the zeta potential data in the Section 3). The final pH of the dispersion was maintained at 8. The mixture was stirred for 20 h to allow complete functionalization of the IO NPs with bPEI. After functionalization, the bPEI-coated iron oxide nanoparticles (IO-bPEI) were separated with a permanent magnet (separation time 30 s), washed several times with MilliQ water, and finally dispersed again in MilliQ water.

2.3. Preparation of IO-TiO₂ Photocatalytic Nanocomposites

A simple heteroagglomeration method utilizing electrostatic interactions between oppositely charged NPs was used to prepare the magnetic photocatalytic IO-TiO₂ nanocomposites, similar to the study by Makovec et al. [45], but using different modification agents and experimental conditions to perform heteroagglomeration. A commercially available dispersion of TiO₂ stabilized with citric acid (i.e., TiO₂@CA) was adjusted to a concentration of 1 mg/mL, with the pH set to 7.5 using 0.1 M NaOH or HCl. An aqueous dispersion containing 200 mg TiO₂@CA NPs in 100 mL was gradually added to the IO-bPEI NPs dispersion. A weight ratio of 1:1 between IO-bPEI and TiO₂@CA NPs was maintained. Similarly, a 0.1 wt% aqueous dispersion of IO-bPEI NPs was prepared at pH 7.5 (200 mL containing 200 mg IO-bPEI). The negatively charged TiO₂@CA NPs were slowly introduced into the positively charged IO-bPEI dispersion with vigorous mixing, and the pH of the mixture was maintained at 7.5 by adding 0.1 M NaOH or HCl as required. The resulting nanocomposite was stirred for 1 h at room temperature to ensure successful heteroagglomeration. The nanocomposite was then separated with a permanent magnet with a time of separation of ~30 s and washed twice with MilliQ water. The overall concept of the work presented here is shown schematically in Scheme 1.

2.4. Characterization of Photocatalytic Nanocomposite

2.4.1. pH Titration Curves

Potentiometric pH titration was used to determine the amount and presumed nature of charged functional groups in aqueous solutions of bPEI and ciprofloxacin. The protonation and deprotonation behavior of the groups, which is pH-dependent, was monitored by potentiometric titrations performed in the forward direction (acidic to alkaline) over the wide pH range. The titrations were performed with HCl (0.1 mol/L) and KOH (0.1 mol/L) as titrants. A T70 multipurpose autotitrator with two burettes (Mettler Toledo, Greifensee, Switzerland) was used to add the titrants, while pH measurements were recorded using an InLab Reach combined glass electrode (Mettler Toledo, Switzerland). Titrations were performed in an inert atmosphere by bubbling nitrogen gas, and an ionic strength of 0.1 mol/L was maintained by adding 3 mol/L KCl. A detailed description of the potentiometric charge titration method can be found in the references [72,73].

2.4.2. X-ray Powder Diffraction (XRD)

The composition of the prepared materials was determined by X-ray powder diffraction (XRD) on a PANanalytical PRO MPD diffractometer with Cu Kα1 radiation (1.54056 Å). Scans were performed over a range of 10–90° in 0.034° increments. Standard reference data was obtained from the International Center for Diffraction Data (PDF).

2.4.3. Transmission Electron Microscopy (TEM)

The morphology and elemental mapping of composites were analyzed using a transmission electron microscope (TEM, JEOL JEM 2100UHR, Tokyo, Japan) operating at 200 kV, equipped with an energy dispersive X-ray spectrometer (EDXS, Oxford X-Max80, Oxfordshire, UK). For TEM observations, the particles were deposited on a carbon-coated copper grid from suspension.

2.4.4. Scanning Electron Microscopy (SEM) and Energy Dispersive X-ray Spectroscopy (EDXS) Mapping

The morphology of IO-bPEI, TiO₂@CA, and IO-TiO₂ was analyzed by scanning electron microscopy (SEM) using a JSM-7610FPlus instrument (Jeol, Tokyo, Japan). About 100 μL of the dispersion was applied to double-sided conductive carbon tape using a pipette, allowed to dry thoroughly, and then mounted on a holder. The sample was sputtered with platinum to further ensure conductivity. SEM images were taken at an accelerating voltage of 10 kV and a magnification of 5000x using a secondary electron detector. In addition, energy dispersive X-ray spectroscopy (EDXS) was performed to identify and map the distribution of certain elements, such as C, N, O, Fe, and Ti.

2.4.5. ATR-FTIR

The infrared spectra of the samples were measured using a Perkin Elmer Spectrum GX NIR FT-Raman spectrometer (Waltham, MA, USA) equipped with a diamond crystal ATR accessory and operating in transmission mode. A background spectrum was recorded first, followed by measurements of the investigated samples under identical conditions. The specific parameters used for the samples included 16 scans with a resolution of 4 cm⁻¹, covering a wavenumber range of 4000–400 cm⁻¹. After scanning, the spectra were processed with automatic smoothing filters, baseline corrections, and ATR corrections and normalized to 1 for comparison purposes. All measurements were performed at room temperature.

2.4.6. X-ray Photoelectron Spectroscopy (XPS)

The chemical composition of the MNP samples was analyzed using X-ray photoelectron spectroscopy (XPS). A PHI-TFA model from Physical Electronics (Chanhassen, MN, USA) was used. The main XPS chamber was maintained at an ultimate pressure of 6 × 10⁻⁸ Pa. The XPS setup included a monochromatic Al Kα1.2 radiation source with a photon energy of 1486.6 eV and a hemispherical analyzer positioned at a 45° angle. The MNP samples were mounted on double-sided copper tape and analyzed over an area with a diameter of 400 µm. An electron gun was used to neutralize the surface charge during the measurements. The C-C component in the C1s peak was calibrated to 248.8 eV. Survey scan spectra were acquired at a pass energy of 187 eV with an energy step of 0.4 eV, while high-resolution spectra were acquired at a pass energy of 27.35 eV with an energy step of 0.125 eV. Data analysis was performed using the Multipak 9.9 software supplied with the spectrometer.

2.4.7. Zeta Potential and Hydrodynamic Diameter

The zeta potential (ZP) and hydrodynamic diameter (HD) of the prepared particle dispersions, formulated as a 0.1 wt% aqueous dispersion in a 1 mM KOH electrolyte background, were determined using a Litesizer 500 analyzer (Anton Paar, Graz, Austria) at 25 °C. The ZP measurements were performed by electrophoretic light scattering (ELS), while the HD measurements were performed by dynamic light scattering (DLS). Prior to analysis, the dispersion was stirred and adjusted to the desired pH using a titration unit with 0.05 M HCl or 0.05 M KOH solutions. The dispersion was then transferred to an Omega cuvette for both ZP and size measurements and placed in the sample holder of the device. The following conditions were used for the HD measurements: pH 8 and at 25 °C, in water as solvent. The selected material was considered unknown due to the different components and complexity of the designed NPs and final nanocomposite. For ZP, the Smoluchowski approximation in water as solvent was used (refractive index 1.3303, viscosity 0.00089 Pas, relative permittivity 78.368). Data acquisition was managed using Kalliope software (version 2.22.1, Anton Paar, Graz, Austria).

2.4.8. Thermogravimetric Analysis and Magnetic Properties

The mass loss as a function of temperature was determined by thermogravimetric analysis (TGA). Approximately 5 mg of each well-dried sample was placed in 60 μL alumina crucibles for analysis using a Mettler TGA/SDTA 851 thermogravimetric analyzer (Mettler Toledo, Switzerland). The samples were heated from 25 to 600 °C at a rate of 10 K/min in a nitrogen atmosphere. The room-temperature magnetization curves of the bare IO NPs and the magnetic IO-TiO₂ nanocomposites were measured using a vibrating-sample magnetometer (VSM, Lake Shore 7307, Westerville, OH, USA). The dried nanoparticle powders were placed in plastic holders and sealed for measurement. The magnetization was recorded at room temperature with an external magnetic field ranging from −10 × 10³ to 10 × 10³ Oe.

2.4.9. Specific Surface Area, Pore Volume, and Average Pore Size of the Materials

To determine the specific surface area, pore volume, and average pore size of the materials, N₂-physisorption experiments were carried out using a Micromeritics TriStar II 3020 device equipped with the SmartPrep degasser. Degassing was performed in an N₂ stream (Linde, Danbury, CT, USA, purity 6.0) at 65 °C for 24 h. Brunauer, Emmett, and Teller (BET) theory was used to calculate the specific surface area of the photocatalysts, while the theory of Barret, Joyner, and Halenda (BJH) was used to determine the pore size distribution.

2.4.10. UV-Vis Diffuse Reflectance (UV-Vis-DR) and Solid Photoluminescence (PL) Spectra

A Perkin-Elmer Lambda 35 UV-Vis spectrophotometer, equipped with an RSA-PE-19M praying mantis powder holder, was used to obtain the UV-Vis diffuse reflectance (UV-Vis-DR) spectra of the synthesized materials. Spectralon© was used for background corrections. All spectra were recorded in absorbance mode from 800 to 250 nm with a scan rate of 240 nm/min and a slit width of 2.0 nm. Kubelka-Munk theory was used to estimate the band gap values.

Solid photoluminescence (PL) spectra were recorded using a Perkin-Elmer LS-55 fluorescence spectrometer equipped with a solid plunger plate and a powder sample holder. The excitation wavelength was set to 300 nm, with a scan rate of 150 nm/min (from 340 to 560 nm) and an emission slit width of 5.5 nm.

2.5. Exploitation of Photocatalytic Performance towards Ciprofloxacin

The photocatalytic performance of the developed nanocomposite with respect to ciprofloxacin was investigated in a 250 mL glass beaker containing 50 mL of an initial 0.008 g/L ciprofloxacin solution (initial concentration in all cases before irradiation). The effect of nanocomposite dosage was investigated (i.e., for 2.5, 5, 10, 15, and 35 mg of IO-TiO₂ nanophotocatalyst), and the effect of pH on the nanocomposite and ciprofloxacin was studied at pH values of 3, 5, 7, and 9. The kinetics of the degradation of ciprofloxacin were investigated for up to 2.5 h (specifically at 30, 60, 90, 120, and 150 min). For conducting the photocatalytic experiments, the photoreactor (Luzchem) was irradiated with artificial UV-B light from 6 overhead UV-B lamps (emitting light in the range of 281–315 nm, with a maximum at λ = 313 nm and an intensity of about 141 W/m²). Prior to UV-B irradiation, the IO-TiO₂ and ciprofloxacin dispersion was stirred in the dark for 30 min to achieve adsorption-desorption equilibrium. Afterwards, the UV-B lamps were activated. After certain irradiation times and examining different effects, the beaker was removed from the photoreactor, and the magnetic nanocomposite was deposited with a permanent magnet. The supernatant was then additionally filtered through a 0.2 μm syringe filter to remove any residual nanocomposite that might interfere with the UV-Vis analysis. The residual ciprofloxacin concentration in the exposed dispersion was determined using a Cary 60 UV-Vis spectrophotometer, and the concentration of remained ciprofloxacin was calculated using a calibration curve (given in Supplementary Information, Figure S1). The absorbance changes of ciprofloxacin under different irradiation conditions were monitored at a maximum wavelength of 277 nm. The photocatalytic degradation was calculated using Equation (1), and the ciprofloxacin degradation efficiency was determined using Equation (2). The reusability of the nanocomposite was evaluated over four cycles, with the recovered nanocatalyst used in subsequent photocatalytic processes under optimal conditions and a constant working volume. Prior to each recycling experiment, the nanocomposite was washed with 0.1 M HCl to remove adsorbed ciprofloxacin, followed by MilliQ water. A constant stirring speed and a fixed distance between the UV-B light source and the solution were maintained throughout the photocatalytic experiments. The changes in the nanocomposite after the reusability cycles were evaluated using ATR-FTIR, thermogravimetric analysis, and surface charge measurements.

Photocatalytic degradation = c/c₀

(1)

Photocatalytic degradation efficiency = c₀ − c/c₀ (%)

(2)

where c₀ is the initial concentration of ciprofloxacin and c is the ciprofloxacin concentration during the investigation of the various influences of degradation.

The kinetics of photocatalytic degradation were investigated using a zero-order kinetic model, a pseudo-first-order kinetic model, and a second-order kinetic model. The description of the individual kinetic models used is given in Supplementary Information.

2.6. Statistical Analysis

Each set of conditions was tested in triplicate to ensure the reliability of our results, and the results are reported as means of three independent experiments, unless otherwise stated. This approach was chosen to minimize experimental variability and to obtain a robust estimate of central tendency. Data analysis was performed using Origin 2018.

3. Results and Discussion

3.1. Characterization of Designed Photocatalytic Nanocomposite

The XRD diffractogram of TiO₂@CA in Figure 1 shows peaks typical of anatase TiO₂ (JCPDS no. 01-075-2550), while the diffractogram of IO-bPEI shows peaks corresponding to γ-Fe₂O₃ (JCPDS no. 04-022-3973). The XRD diffractogram of IO-TiO₂ shows peaks for both anatase TiO₂ and Fe₂O₃. The average crystallite size of anatase in the TiO₂@CA and IO-TiO₂ samples was calculated from the width of the (001) diffraction peak (2θ = 25°) using the Scherrer formula. The results presented in Table S1 show that the average crystallite size of anatase in IO-TiO₂ remained unchanged during the synthesis, indicating that IO-bPEI was deposited only on the surface of TiO₂@CA, which is in line with the conducted hetero-agglomeration method.

Transmission electron microscopy (TEM) was used to investigate the morphology and microstructure of heteroagglomerated IO-TiO₂ nanocomposites (Figure 2) as well as single IO NPs (Figure S2), and TiO₂@CA NPs (Figure S3). The TEM image of IO-bPEI nanoparticles (Figure S2a) shows quasi-spherical nanoparticles with a large polydispersity. The Selected Area Electron Diffraction (SAED) pattern (Figure S2b) showed the polycrystalline nature of the nanoparticles, with the diffraction rings matching well with the X-ray diffraction (XRD) data (Figure 1), suggesting a spinel crystal structure for the IO-bPEI nanoparticles. High-resolution TEM images (Figure S2c) confirmed the crystallinity of the prepared nanoparticles, which had an average size of about 13 nm. In contrast, the TiO₂@CA NPs (Figure S3a) exhibited a significantly different morphology than IO-bPEI. These NPs were smaller, with a size of about 3–5 nm, and also exhibited good crystallinity (Figure S3c). The SAED pattern (Figure S3b) was consistent with the XRD results (Figure 1), indicating an anatase crystal structure for the TiO₂@CA NPs. When both types of NPs were combined in a composite, the resulting nanocomposite showed integration of both nanoparticle types, with agglomerates of IO-bPEI often covered by TiO₂@CA and vice versa (Figure 2). The size of the nanocomposite ranged from hundreds of nanometers to the micrometer range (Figure 2a,b), which was consistent with the measurements of the hydrodynamic diameter of the nanocomposite (results below). In agreement with the XRD pattern, the nanocomposite contained both types of NPs, as confirmed by energy dispersive X-ray spectroscopy (EDXS) in STEM mode, which showed the presence of elements characteristic of both types of nanoparticles (Figure 2c–e). Importantly, it was demonstrated that the magnetic removal of NPs is significantly more effective when the magnetic nanoparticles (MNPs) form clusters [74]. This finding underlines the crucial role of the nanocomposite structure in enhancing the efficiency of magnetic separation [45].

The morphological features of the designed photocatalytic nanocomposite were investigated by scanning electron microscopy (SEM) to observe the formed structure at lower magnification (Figure 3a, Figures S4 and S5). Elemental mapping of the individual components of the nanocomposite was also performed (Figure 3b–g), together with EDXS spectra analysis (Figure 3h). The same characterization was performed for IO-bPEI (Figure S4) and TiO₂@CA (Figure S5). The SEM images revealed that the formed IO-TiO₂ nanocomposite has a rather homogeneous structure, which is advantageous for the improvement of the photocatalytic process. The elements present in IO-bPEI and TiO₂@CA were also confirmed by elemental mapping (Figures S4 and S5). However, due to the limited magnification of the SEM images, individual NPs and the detailed structure of IO-TiO₂ could not be clearly recognized, in contrast to Figure 2 and Figures S2 and S3 (Supplementary Information). The TEM was used for this purpose (Figure 2). Elemental mapping of the typical elements in the nanocomposite (C, N, O, Fe, and Ti; Figure 3b–g) showed a uniform distribution, confirming the successful formation of the nanocomposite with moderately uniformly distributed NPs. The EDXS spectrum (Figure 3h) also confirmed the presence of all constituent elements in the IO-TiO₂ nanocomposite.

The results of ATR-IR spectroscopy for all samples are shown in Figure 4. The peak at 534 cm⁻¹ associated with Fe-O (in IO NPs), as identified in [75], is present in both the IO-bPEI and IO-TiO₂ samples, indicating the successful formation of the designed nanocomposite. The peaks at 1620 cm⁻¹ and around 3200 cm⁻¹ correspond to the O-H bending and stretching of the water molecules in the IO NPs. The successful functionalization of IO NPs with bPEI is evidenced by new peaks. In particular, the peaks at 3400 cm⁻¹ and 1628 cm⁻¹ correspond to N-H stretching and bending of primary amines, while the peak at 1457 cm⁻¹ corresponds to N-H stretching of secondary amine groups [76]. In addition, the C-N stretching peak at 1315 cm⁻¹ confirms the presence of amine-containing molecules [77]. The presence of TiO₂ is confirmed by typical peaks at 804 cm⁻¹ and 616 cm⁻¹, which are attributed to Ti-O stretching [78]. In addition, the presence of citric acid on TiO₂ NPs was suggested by peaks characteristic of carboxylic acids. This was evident from the peaks at 1633 cm⁻¹ and 1385 cm⁻¹, which correspond to the symmetric and asymmetric stretching of the C=O group [49], respectively. After hetero-agglomeration of TiO₂@CA with IO-bPEI, these peaks are less pronounced, which is probably due to a lower concentration. However, the broadening of the peak indicates the formation of hydrogen bonds [79] and, furthermore, an electrostatic interaction between negatively charged citric acid and positive bPEI. Indeed, the peaks present on the individual NPs are also clearly visible in the nanocomposite IO-TiO₂ (Figure 4). Thus, infrared spectroscopy confirmed the successful functionalization of the IO-bPEI particles with bPEI and the hetero-agglomeration of TiO₂@CA with IO-bPEI.

FTIR results are further supported by XPS results.

Table 1. Elemental surface composition in at % for different samples.

Sample/Element (at %)	C	O	Fe	N	Ti
IO NPs	14.7 ± 1.0	65.8 ± 1.8	19.5 ± 1.3
IO-bPEI	18.6 ± 1.1	57.3 ± 0.7	17.8 ± 1.1	6.3 ± 0.7
Citric acid (CA)	49.0 ± 1.6	51.0 ± 1.4
TiO2@CA	12.4 ± 2.0	62.3 ± 1.4			25.3 ± 1.1
IO-TiO2	21.9 ± 2.2	56.2 ± 1.0	7.5 ± 0.6	2.9 ± 1.2	11.6 ± 1.1

shows the XPS elemental composition of the NPs. Additionally, we also show in Figure 5a recorded XPS survey spectra of the samples IO-bPEI, TiO₂@CA, and IO-TiO₂, with characteristic peaks used for obtaining the elemental composition shown in Table 1. For the sample IO NPs, carbon, oxygen, and iron were found, which is consistent with their expected composition. The presence of a layer of bPEI (sample IO-bPEI) is confirmed by the obvious presence of nitrogen originating from amino groups in bPEI. In the case of TiO₂ samples coated with CA, the presence of large amounts of oxygen and substantial amounts of titanium, as well as carbon, was found. Oxygen originates from both TiO₂ and CA. The composition of the reference CA sample is also shown in Table 1, and we can see that oxygen concentration prevails over carbon, as found also for the TiO₂@CA sample. For the final composite IO-TiO₂ samples, we found all the elements that are characteristic of the individual components from which they are composed, i.e., iron, nitrogen, titanium, carbon, and oxygen. This result clearly demonstrates the success of the nanocomposite formation. High-resolution spectra of Fe, Ti, N, C, and O were measured as well. They are shown in Figure 5b–f. Iron peak Fe2p_3/2 was found at a binding energy of 710.9 eV (Figure 5b). There is also a small satellite peak at approximately 719.5 eV, which is characteristic of Fe³⁺ iron [80]. The satellite peak is not as noticeable on the IO-TiO₂ sample as on IO-bPEI, which could be explained by the coverage of the NPs surface with TiO₂@CA particles. The titanium Ti 2p peak of TiO₂@CA samples was observed at a binding energy of 458.9 eV, which is typical for TiO₂ oxide (Figure 5c) [81,82]. For IO-TiO₂, we observe, in addition to the peak characteristic of TiO₂, another peak at a higher binding energy of approximately 460 eV. There is no data for this Ti 2p peak at such a high energy in the literature, so it can only be attributed to the presence of IO, which affects the chemical environment of titanium and thus causes changes in binding energies due to mutual interactions between TiO₂@CA and IO-bPEI. In Figure 5d,e we show nitrogen and carbon peaks, which are related to the presence of organic molecules. As already mentioned, the origin of nitrogen is the amino groups of bPEI. Therefore, nitrogen is observed in the IO-bPEI sample and in IO-TiO₂, where its concentration is lower (Table 1) because of the presence of TiO₂@CA particles on the NPs surface and vice versa. A comparison of carbon peaks in Figure 5e and fitted carbon spectra in Figure 5f can provide information on the presence of CA because CA is characterized by carboxyl groups. In Figure 5f, we observe in the C1s spectrum of TiO₂@CA the peak related to COOH groups, which clearly confirms the presence of CA on the surface. To conclude, the XPS results presented below prove the successful hetero-agglomeration of nanocomposite particles.

At low pH (acidic conditions, e.g., 3–4), the surface of magnetic nanoparticles (IO NPs) is usually positively charged due to the adsorption of hydrogen ions (H⁺) and the zeta potential might be around +30 mV (Figure 6a). Above pH 5, it starts to decrease until reaching the isoelectric point (e.g., pH 5.8), where the zeta potential approaches 0 mV. At high pH (above the isoelectric point), the surface is typically negatively charged due to the adsorption of hydroxide ions (OH⁻). It reaches a negative zeta potential plateau value at around pH 7.5. Positive and negative zeta potential plateau levels are reached at the absolute value of zeta potential, where the stability of colloidal dispersion is guaranteed [83]. Interestingly, TiO₂@CA shows very similar behavior of zeta potential as a function of pH as bare IO NPs. The zeta potential of TiO₂@CA nanoparticles as a function of pH exhibits similar characteristics to IO NPs, which is reasonable since both are metal oxide nanoparticles with a tendency to be positive in acidic conditions due to the adsorption of H⁺ and negative at higher pH due to the adsorption of hydroxide ions (OH⁻) [75]. Both zeta potential plateau levels, positive and negative, are slightly higher (positive) and lower (negative) in comparison to IO NPs, which may be connected to CA, especially by the lower negative zeta potential plateau level where carboxylic acids of CA are available on the particles surface to be in protonated form [49]. IO-bPEI shows successful functionalization with bPEI, where the zeta potential is positive across the entire pH range due to the protonated amino groups [76], in line with the potentiometric titration curve of bPEI (Figure S6), where the highly positive character of the latter is revealed. There is a tendency of decreasing positive zeta potential slowly above pH 7, which may be due to less accessibility of amino groups due to their deprotonation (see potentiometric titration curve of bPEI) and thus to possible slight agglomeration. IO-TiO₂ nanocomposites show the dominating character of IO-bPEI due to almost the same ZP (pH) curve as for IO-bPEI. This means that at hetero-agglomeration of IO-TiO₂ nanocomposites by electrostatic interactions among IO-bPEI and TiO₂@CA NPs, IO-bPEI chemistry dominates.

Successful formation of nanocomposites at pH 7.5 is evident through the increase in particle size (Figure 6b) from the colloidal range of the individual systems (TiO₂@CA and IO-bPEI, respectively) to the micron range when combined in the composite (IO-TiO₂). In this context, the IO-bPEI nanoparticles exhibit a relatively narrow size distribution, with an average diameter of about 200 nm. This narrow distribution can be attributed to the positive repulsive forces, as shown by the zeta potential (ZP) of IO-bPEI, which is ~25 mV at pH 8. The TiO₂@CA nanoparticles also have an average size of about 200 nm but exhibit a broader size distribution compared to IO-bPEI. Although both nanoparticles have similar absolute ZP values (about 25 mV at pH 8), TiO₂@CA NPs tend to form larger agglomerates, which is consistent with the TEM analysis (Figure S3). As expected, the hydrodynamic diameter of the nanocomposite is larger than that of the individual components and is about 2260 nm, which is in good agreement with the TEM observations (Figure 2). Although the highest ZP value at pH 8 was measured for the IO-TiO₂ nanoparticles, which typically indicates smaller particle sizes due to stronger repulsive forces, the attractive interactions between the individual nanoparticles seem to dominate. This leads to the formation of a nanocomposite with a size in the micrometer range and a moderately uniform size distribution.

Thermogravimetric analysis (TGA) was performed to determine the mass loss of the different nanomaterials, and the resulting curves were compared (Figure 7a). The bare IO NPs showed a minimum mass loss of about 3–4 wt%, probably due to the presence of physically adsorbed water. After modification with bPEI, the IO NPs showed a more significant mass loss of about 5 wt%, indicating the presence of an organic coating on the IO-bPEI. For the TiO₂@CA nanoparticles, the TGA curves indicated a mass loss of 12 wt%, which is likely due to the citric acid (CA) coating and reveals distinct decomposition patterns. When TiO₂@CA was combined with IO-bPEI, the nanocomposite displayed a maximum mass loss of 11 wt% at 600 °C. This indicates a successful formation of the nanocomposite. The IO-TiO₂ nanocomposite exhibited a TGA curve that reflects the properties of both nanomaterials of which it is composed. The organic species within the nanocomposite began to decompose above 120 °C and decomposed almost completely up to 600 °C, likely resulting in volatile products [84]. This thermal behavior is further evidence of the successful integration of the components into the nanocomposite. The magnetic properties of the bare IO NPs and the synthesized IO-TiO₂ nanocomposite were investigated by hysteresis loop measurements. As illustrated in Figure 7b, both the bare IO NPs and the nanocomposite exhibit superparamagnetic behavior characterized by negligible coercivity and remanence [85,86]. This is of crucial importance since superparamagnetic NPs do not exhibit spontaneous magnetic moments and thus prevent agglomeration due to magnetic dipole-dipole interactions, in contrast to larger particles [87,88]. The saturation magnetization of the IO NPs decreased from an initial 67 emu/g to 49 emu/g when combined TiO₂@CA and IO-bPEI. This decrease is due to the presence of organic coatings and non-magnetic material within the nanocomposite. Despite this decrease, the IO-TiO₂ nanocomposite retained remarkable magnetic properties that enabled effective magnetic separation of the nanocatalyst after the photocatalytic process in the presence of an external magnetic field. The strong magnetic response of the nanocomposite is significant as it enables not only enhanced photocatalytic activity but also efficient removal and reusability, minimizing additional environmental pollution. This dual functionality underlines the potential of the developed nanocomposite for practical applications in environmental remediation.

Since the specific surface area (S_BET) has a significant influence on the photocatalytic activity of a material [89,90], whereby a catalyst with a lower S_BET value generally has a lower photocatalytic activity than a catalyst with a higher S_BET value, nitrogen adsorption/desorption experiments were carried out to analyze the porosity of the materials investigated (Figure 8). The specific surface areas (S_BET) were calculated using the Brunauer-Emmett-Teller (BET) theory. The isotherms for IO-bPEI, TiO₂@CA, and IO-TiO₂ (Figure 8a) correspond to the IUPAC classification of type IV with H1 hysteresis loops, indicating a mesoporous structure with cylindrical pores [91,92]. As shown in

Table 2. Comparison of the specific surface area (S_BET), average pore diameter (d_pore), and total pore volume (V_pore).

Sample	SBET	Vpore	dpore
	m2/g	cm3/g	nm
IO-bPEI	55.8	0.24	14.5
TiO2@CA	246.8	0.47	7.6
IO-TiO2	165.3	0.45	10.1

, IO-bPEI has an S_BET value of 55.8 m²/g, which is significantly lower than the S_BET value of TiO₂@CA of 246.8 m²/g. The low S_BET value of IO-bPEI could have a negative impact on its photocatalytic activity compared to TiO₂@CA, as it provides less surface area for light-generated charge carriers to generate reactive oxygen species or directly degrade pollutants in water. In addition, the adsorption of degradation products on the catalyst surface during heterogeneous photocatalysis would affect the catalytic activity of IO-bPEI more than that of TiO₂@CA, which has a higher S_BET. When IO-bPEI nanoparticles were combined with TiO₂@CA, a reduction in specific surface area (S_BET) was observed compared to TiO₂@CA alone. However, the specific surface area remained significant at 165.3 m²/g, which is advantageous for the adsorption and photodegradation of ciprofloxacin (Table 2). Despite the smaller pore volume (V_pore) of IO-bPEI, the V_pore of the nanocomposite increased to 0.45 cm³/g, comparable to that of TiO₂@CA, indicating the presence of mesopores. A similar trend was observed for the pore diameter (d_pore) for IO-TiO₂, with a value of 10.1 nm, which falls between the values for IO-bPEI and TiO₂@CA (Figure 8b, Table 2). The presence of large mesopores and a large effective surface area likely facilitates charge separation and increases the number of active sites within the nanocomposite structure, improving its suitability for heterogeneous photocatalytic applications.

3.2. Photocatalytic Degradation of Ciprofloxacin

During photocatalytic degradation of antibiotics with TiO₂-based magnetic nanocomposites, photons from the light source excite electrons from the valence band to the conduction band, creating electron-hole pairs. These electron-hole pairs initiate the degradation process by generating various radicals through reactions with water molecules. The generated radicals then degrade the antibiotic molecules adsorbed on the nanocomposite surface through oxidative and radical-driven reactions, transforming the antibiotics into less harmful by-products [20,61]. In particular, the magnetic properties of the iron oxide NPs allow easy manipulation and removal of the nanocomposite by an external magnetic force, so that it can be reused in several photodegradation cycles. Remarkably, the iron oxide NPs can also help to improve photocatalytic performance. The UV-Vis reflectance spectra (UV-Vis-DR) and the corresponding Kubelka-Munk functions for IO-bPEI, TiO₂@CA, and IO-TiO₂ are shown in Figure 9a and Figure S7. The band gap values of these solids determined using the Kubelka-Munk function are listed in

Table 3. Comparison of the band gap values of the investigated materials obtained with the Kubelka-Munk function (Supplementary Information Figure S7).

Sample	Band Gap
	eV
IO-bPEI	1.7
TiO2@CA	3.2
IO-TiO2	1.8

. The TiO₂@CA sample exhibits strong light absorption below 400 nm with a band gap of 3.2 eV, which is typical for anatase TiO₂, [93]. This is consistent with the XRD analysis, which shows peaks in the XRD diffractogram of TiO₂@CA that correspond to anatase TiO₂. The UV-Vis-DR spectra of IO-bPEI show that this sample can absorb light above 400 nm with a band gap of 1.7 eV (typical for iron oxide NPs [94]), which makes IO-bPEI photocatalytically active when illuminated with visible light. In the UV-Vis-DR spectra of IO-TiO₂, the contributions of TiO₂@CA below 400 nm and of IO-bPEI above 400 nm are obvious. The UV-Vis-DR measurements show that the addition of IO-bPEI to TiO₂@CA significantly reduces the band gap of TiO₂ and that both components of the IO-TiO₂ composite can generate charge carriers when illuminated with UV-B light. A key parameter that determines the catalytic activity of a photocatalyst is the recombination rate of the charge carriers generated by the light; a lower recombination rate is associated with a higher photocatalytic activity. The recombination of charge carriers generates energy in the form of photoluminescence (PL) emissions, i.e., the intensity of the PL signal of a solid material is proportional to its electron-hole recombination rate [29,95]. The PL spectra of TiO₂@CA and IO-TiO₂ in the solid state are shown in Figure 9b. TiO₂@CA shows well-defined peaks at about 3.2, 2.9, 2.7, 2.5, and 2.3 eV. The peak at 2.9 eV corresponds to the lowest indirect transition Γ_1b → X_1a, while the signals around 3.1 eV and above are assigned to direct (X_1b → X_2b and X_1b → X_1a) and indirect (X_1b → Γ₃ and Γ_1b → X_2b) transitions. The peaks at 2.8, 2.7, 2.55, and 2.34 eV are associated with shallow traps in anatase [96]. The PL emission intensity of IO-TiO₂ is lower than that of TiO₂@CA, indicating a reduced recombination rate of electron-hole pairs. Similar findings were observed in the study of Sakineh et al. [35], who investigated a composite of TiO₂ decorated on ferroferric oxides coupled with activated carbon (TFOC) under UV light for tetracycline degradation. The photocatalytic performance of the composite was enhanced by the presence of ferroferric oxides, which played a crucial role in suppressing the recombination of excited electrons and holes. In addition, another study showed that the combination of TiO₂ with superparamagnetic iron oxide nanoparticles (SPIONs) significantly enhanced the degradation of dyes. Here, the inclusion of iron oxides extended the adsorption edges of the nanocomposite and improved its overall photocatalytic performance [97].

3.2.1. Influence of Nanocomposite Dosage

The initial parameter that significantly influences the efficiency of the photocatalytic process is the dosage of the IO-TiO₂ photocatalyst. In this study, dosages from 2.5 mg to 25 mg IO-TiO₂ were investigated (Figure 10a, UV-Vis spectra in Figure S8). The photocatalytic activity increased significantly with the dosage of the nanocomposite up to 10 mg. Beyond this dosage, however, no further significant improvements were observed, so that a degradation efficiency of 75–80% was achieved. The initial increase in photocatalytic degradation of ciprofloxacin can be attributed to the increased quantity of nanocomposite, which in turn increases the total effective surface area. This leads to more redox-active centers on the surface of the nanocatalyst and the formation of more radicals or active oxidants responsible for degradation [34,35]. However, a further increase in the IO-TiO₂ dosage led to increased turbidity of the dispersion, which in turn reduced the penetration of UV-B light. This reduction in light penetration limits the number of photons reaching the surface of the nanocatalyst and the excitation of the water molecules, thereby reducing the formation of hydroxyl radicals [34]. Another possible explanation for the reduced photocatalytic performance at higher nanocomposite dosages is agglomeration, which reduces the effective surface area during photocatalytic degradation and thus significantly affects the overall degradation efficiency [35]. Consequently, the optimal IO-TiO₂ dosage was set at 10 mg, resulting in a degradation efficiency of approximately 65%. At higher dosages, the degradation efficiency did not differ significantly and ranged between approximately 70% and 80% (Figure 10a). Therefore, 10 mg of the photocatalyst was used in the subsequent studies to investigate further effects on the photocatalytic degradation of ciprofloxacin.

3.2.2. Influence of pH

One of the most important parameters influencing the photocatalytic degradation of ciprofloxacin is the pH of the dispersion during irradiation. The pH not only influences the surface chemistry of the nanocatalyst but also the charging behavior of the ciprofloxacin molecules. This in turn affects the interaction between the surface of the nanocatalyst, the ionized forms of ciprofloxacin, and the formation of hydroxyl radicals. Adsorption plays a decisive role in the photocatalytic degradation process. The potentiometric titration of pure ciprofloxacin (Figure S9) shows three pK_a values up to a pH of 10, namely pK_a1 = 2.8, pK_a2 = 5.9, and pK_a3 = 9.3, indicating that ciprofloxacin is in a cationic form below pH 6 (see the potentiometric curve in Supplementary Information, Figure S9) and in an anionic form above pH 9.3. Similarly, zeta potential measurements of IO-TiO₂ show that the nanocomposite is positively charged, with the charge decreasing above pH 9 (Figure 10b). In our system, the low photocatalytic degradation at acidic pH (about 15% at pH 3) can be attributed to the repulsive forces between the positively charged IO-TiO₂ (primary amines) and the positively charged ciprofloxacin molecules (due to the protonation of the piperazine ring nitrogen and secondary amines), which reduce the adsorption on the nanocatalyst surface (Figure 10b; UV-Vis spectra are presented in Figure S10). As the pH value increases, photocatalytic degradation also increases. At a pH of 7, there is an electrostatic attraction between the zwitterionic ciprofloxacin molecules (carboxylic acids) and the positively charged IO-TiO₂ (secondary and tertiary amines of bPEI dominating at its surface), resulting in a photocatalytic degradation efficiency of 62%. A further increase in pH to 9 leads to a degradation efficiency of about 70%, which is probably due to electrostatic attractions between the deprotonated carboxylic acids of ciprofloxacin and the tertiary amines of -bPEI as dominating in IO-TiO₂, as well as enhanced van der Waals interactions between the negatively charged ciprofloxacin and the positively charged IO-TiO₂ nanocomposites, as reported in the literature [20,34]. Under alkaline conditions, the aqueous dispersion is more favorable for the efficient formation of OH^• and O₂^−• radicals, leading to higher photodegradation efficiency. In particular, a higher concentration of OH- under alkaline conditions leads increasingly to oxidative OH^• species, which enhance the degradation of ciprofloxacin. Since the degradation efficiency is comparable at neutral and alkaline pHs, neutral pH is preferred in view of practical applications. This minimizes the need for extensive pH adjustment and the consumption of acids and bases, making the process more feasible and cost-effective for practical applications.

3.2.3. Effect of Nanocatalyst Type and Influence of Time on Ciprofloxacin Degradation

To evaluate the effects of the individual components of IO-TiO₂ on the degradation of ciprofloxacin, both IO-bPEI and TiO₂@CA were examined over a period of 150 min under identical conditions (Figure 11a). IO-bPEI showed almost negligible photocatalytic degradation of ciprofloxacin under UV-B light irradiation, which can be attributed to the smaller band gap of maghemite (approx. 1.7 eV, Table 3). The degradation efficiency of the magnetic nanoparticles remained relatively constant and fluctuated between 15–30% over time. In contrast, TiO₂@CA nanoparticles showed a significantly different behavior compared to IO-bPEI. Under the same conditions, TiO₂@CA achieved ciprofloxacin degradation of up to 80% after 150 min of irradiation, showing a consistent degradation trend with prolonged UV-B light exposure. It should be noted that the specific surface area of IO-bPEI was about 4.5 times smaller than that of TiO₂@CA (Figure 8, Table 2), which may also affect the photocatalytic activity. This high activity of TiO₂ semiconductor NPs in generating reactive oxygen species is well documented [51] and leads to improved photocatalytic performance. Interestingly, when comparing the degradation efficiency of the nanocomposite with the individual nanoparticles, some synergistic effect can be observed. Yet, the degradation curve and efficiency of the nanocomposite matched those of TiO₂@CA to some extent, suggesting that TiO₂@CA primarily affects the overall photocatalytic degradation efficiency of ciprofloxacin in the IO-TiO₂ composite. In particular, the degradation of ciprofloxacin reached about 70% after 150 min (Figure 10a). The change in the height of the peak at 277 nm with longer degradation time is shown in Figure 11b. Comparison of different kinetic models (Figure S11 in the Supplementary Information) showed moderate fitting parameters, with IO-TiO₂ being the best fit to a pseudo-first order kinetic model (R² = 0.7766). The rate constants of degradation also varied between the three types of nanoparticles (Figure 11c). The highest rate constant derived from the pseudo-first-order model was observed for TiO₂@CA at 0.0083 min⁻¹, while IO-TiO₂ had a slightly lower rate constant of 0.0062 min⁻¹.

3.2.4. Reusability and Stability Evaluation of the Nanocatalyst

A key factor for practical applications of photocatalysts is their reusability. Reusability minimizes costs as the nanocatalyst can be regenerated and reused in multiple photocatalytic processes, contributing to the development of environmentally friendly and sustainable photocatalysts. To investigate the reusability of IO-TiO₂ for the degradation of ciprofloxacin, four cycles were performed, the results of which are shown in Figure 12. After each process, the IO-TiO₂ was magnetically separated from the dispersion, regenerated with acid, and then reused in the next photocatalytic cycle. The photocatalytic performance remained stable in the first two cycles. However, in the third cycle, a decrease in efficiency of about 15% was observed after 60 min of irradiation. In the fourth cycle, the photocatalytic efficiency had dropped significantly to 40% after 60 min, which is less than half of the efficiency observed in the first cycle. This decrease can be attributed to possible leaching of Ti from the nanocomposite during the acidic wash process or to the blocking of active sites on the photocatalyst by by-products formed in previous cycles that could not be completely removed or by adsorbed ciprofloxacin residues [35]. Despite the decrease in performance after several cycles, the photocatalytic efficiency of IO-TiO₂ did not decrease significantly after three cycles, indicating its potential for practical environmental and economic applications. Structural analysis by infrared spectroscopy after the photocatalytic tests revealed no significant changes in the position of the infrared bands (Figure 13a). Thermogravimetric analysis indicated a slight mass loss (Figure 13b), probably due to desorption of the organic coatings, and a shift of zeta potential towards a more acidic range (Figure 13c), possibly related to the adsorption of more acidic surface-charged compounds after photocatalysis, such as remained ciprofloxacin (Figure S9, potentiometric titration curve) or possibly adsorbed residue degradation products. In addition,

Table 4. Comparison of the photocatalytic performance of different materials for the degradation of ciprofloxacin in the existing literature under different experimental conditions.

Studied Catalyst	Photocatalytic Conditions	Ratio of Catalyst/Ciprofloxacin	Illumination Source	Degradation Efficiency of Ciprofloxacin at (Duration)	Reference
Chlorophyll sensitized and salicylic acid functionalized TiO2 nanoparticles	0.75 g/L catalyst dosage, initial pH of 6, and 10 ppm of initial ciprofloxacin	75	with visible light (λmax = 457 nm)	~75% (120 min)	[58]
Fe2O3 nanoparticles loaded on graphitic carbon nitride	pH of 7, photocatalytic dosage of 0.3 g/L, and 25 mg/L of ciprofloxacin	12	under visible light radiation	100% (60 min)	[59]
defective TiO2-x nanomaterial	10 ppm CIP solution in the presence of 0.5 g/L TiO2	50	the visible range	about 10% (180 min)	[60]
magnetite (Fe3O4) nanoparticles	pH 6 for ciprofloxacin, dosage of the photocatalyst (0.12 g), concentration (100 mg/L),	1.2	sunlight radiation	73.51% for magnetic nanocomposites, 54.57% magnetic nanoparticles (240 min)	[47]
waste polystyrene and TiO2 composites	a pH value of 3 and an initial ciprofloxacin concentration of 5 mg/L	/	(400–800 nm) used as a light source	95.01% (180 min)	[34]
TiO2 nanoparticles immobilized on a glass plate	pH of 5, ciprofloxacin initial concentration of 3 mg/L.	/	UV-C irradiation	86.57% (105 min)	[33]
Fe3O4/P(NIPAM-co-MAA)/Ag-TiO2	initial concentration of 5 mg/L; concentration of the nanocomposites: 460 mg/L	92	visible-light irradiation	47%	[54]
γ-Fe2O3/Bi2WO6	0.1 g photocatalysts and 100 mL ciprofloxacin (15 mg/L)	67	visible light irradiation	65% (120 min)	[98]
ZnIn2S4/CoFe2O4 p-n junction-decorated biochar	50 mL ciprofloxacin solution (20 mg/L, pH = 7)	25	simulated sunlight	96.9% (120 min)	[63]
CuFe2O4@methyl cellulose based magnetic nanobiocomposite	pH = 7, ciprofloxacin initial concentration of 3 mg/L, photocatalyst loading of 0.2 g	/	UV-C lamps	80.74% (90 min)	[66]
Ni-doped iron oxide/graphitic carbon nitride nanocomposites	catalyst = 15 mg in 50 mL of ciprofloxacin with 10 mg/L,	30	solar light exposure	82.1% (120 min)	[68]
IO-TiO2 heterostructures	optimal pH of 7, and ciprofloxacin initial concentration of 0.008 g/L and 50 mL, 10 mg of nanocomposite	25	UV-B irradiation	Approx. 70% (150 min)	This study

shows a comparison between the results of this study and more recent research dealing with the photocatalytic degradation of ciprofloxacin. Notably, the recyclable and magnetically controlled nanocomposite employed in this research demonstrated satisfactory performance.

3.3. Plausible Degradation Mechanism

A promising strategy to improve the photocatalytic efficiency between two different semiconductors is heterojunction coupling [30], as represented in the current system by the combination of TiO₂@CA and IO-bPEI nanoparticles. This hybrid configuration offers several advantages for photocatalysis, including improved charge separation, extended carrier lifetime, increased recombination resistance, and more efficient charge transfer at the interface to the adsorbed molecules [30]. According to photoluminescence (PL) measurements in the UV-B spectral range, both the individual nanoparticles and the IO-TiO₂ nanocomposite can generate charge carriers that migrate between the phases and thus reduce recombination rate (Figure 9b). Although the nanocomposite consists of a 1:1 ratio of TiO₂@CA to IO-bPEI, its overall photocatalytic performance is almost identical to that of pure TiO₂@CA under optimal conditions, emphasizing the significant role of IO-bPEI nanoparticles. The presence of IO-bPEI likely supports the phase separation between TiO₂@CA and IO-bPEI [99], which increases the photocatalytic efficiency. XPS analysis was used to determine the positions of the valence bands (VB) of TiO₂@CA and IO-bPEI by measuring their valence band maxima (VBM). The data presented in Figure S12 show that the VBM values for TiO₂@CA and IO-bPEI are 2.5 eV and 0.30 eV, respectively. When we combine these VBM results with the band gap values obtained from UV-Vis diffuse reflectance (UV-Vis DR) measurements, we can create a detailed schematic representation of the energy band structure for IO-TiO₂ and propose the mechanism of charge transfer under UV-B light (Figure S13). When irradiated with UV-B light, both TiO₂@CA and IO-bPEI generate charge carriers. Holes in the VB of TiO₂@CA can migrate into the VB of IO-bPEI, while electrons from the conduction band (CB) of IO-bPEI can migrate into the CB of TiO₂@CA. This process promotes the separation of electron-hole pairs, increases their concentration on the surface of the IO-TiO₂ nanocomposites, and thereby enhances the overall photocatalytic activity. This phenomenon is confirmed by PL measurements, which show a significantly lower PL intensity for the composite material compared to TiO₂@CA alone, indicating a more efficient charge separation and extended carrier lifetime in the composite material. Consequently, this leads to the formation of reactive oxygen species (ROS) such as O₂^−• and OH^•, which are highly reactive and can react effectively with ciprofloxacin to form less harmful degradation products (Figure 14) [36,45,48]. In contrast to our findings, it was observed that when the α-Fe₂O₃@TiO₂ system was irradiated with UV light, the TiO₂ was excited as expected, but the α-Fe₂O₃ acted as a recombination center for photoinduced charge carriers [56]. Consequently, the designed composite showed relatively poor photocatalytic activity. One possible reason for this discrepancy could be the different iron oxide used in the composite. In our study, maghemite with a spinel crystal structure was used as the magnetic nanoparticles, while [56] used hematite as the iron oxide.

As already mentioned, the photocatalytic degradation of ciprofloxacin with IO-TiO₂ under the influence of light and excitation by photon energy (with an energy equal to or greater than the band gap of the photocatalyst) leads to the formation of electron (e⁻)—hole (h⁺) pairs (h⁺ in valence band VB and e⁻ in conductive band CB). These pairs subsequently generate highly reactive species such as superoxide anions (O₂^−•) and hydroxyl radicals (OH^•). The proposed degradation mechanism caused by photocatalytic nanocomposite is suggested by Equations (3)–(7). Usually mentioned reactive species primarily attack the C-N bond of the piperazine ring, the fluorine atom, the carbonyl group, and the cyclopropyl group within the quinolone structure, leading to oxidation and degradation of the ciprofloxacin molecules, as reported in the current study [20]. A similar photodegradation pathway for ciprofloxacin, driven by O₂^−• and OH^• radicals through the generation of electron-hole (e⁻/h⁺) pairs, has been observed in other photocatalytic systems. These include chlorophyll-sensitized and salicylic acid-functionalized TiO₂ nanoparticles [58], Fe₂O₃ nanoparticles supported on graphitic carbon nitride [59], magnetic nanocomposites [47], composites of polystyrene waste with TiO₂ [34], and TiO₂ nanoparticles immobilized on glass substrates [33].

IO-TiO₂ + hν (UV-B light) → IO-TiO₂ (e_CB⁻ + h_VB⁺)

(3)

IO-TiO₂ (e_CB⁻) + O₂ → IO-TiO₂ + O₂^−•

(4)

IO-TiO₂ (h_VB⁺) + H₂O/OH⁻ → IO-TiO₂ + OH^•

(5)

O₂^−• + ciprofloxacin molecules → ciprofloxacin degradation products

(6)

OH^• + ciprofloxacin molecules → ciprofloxacin degradation products

(7)

The heterogeneous photocatalytic process described in this report comprises the following steps:

(i)

Transfer of ciprofloxacin from the aqueous dispersion to the surface of the IO-TiO₂ nanocomposite.

(ii)

Adsorption of ciprofloxacin onto IO-TiO₂.

(iii)

Photocatalytic oxidation and reduction of ciprofloxacin by reactive species as described in Equations (3)–(7).

(iv)

Removal of ciprofloxacin degradation products from the IO-TiO₂ surface.

(v)

Transport of these degradation products from the IO-TiO₂ surface back into the aqueous dispersion.

(vi)

Magnetically assisted removal of IO-TiO₂ after the photocatalytic process.

(vii)

Reuse of IO-TiO₂ for another photocatalytic process.

4. Conclusions

The world is currently struggling with significant residues of antibiotics entering aqueous systems, with ciprofloxacin being a major contributor. This is exacerbating the alarming rise in antimicrobial resistance (AMR). In this pioneering study, we successfully synthesized a magnetically controlled and photocatalytic nanocomposite IO-TiO₂ for the first time by hetero-agglomerating positively charged bPEI-modified iron oxide (IO) NPs with commercially available negatively charged TiO₂@CA NPs. This novel nanocomposite was tested for its ability to degrade ciprofloxacin in aqueous media under UV-B irradiation, an investigation not previously reported in the literature. The TiO₂@CA exhibited an anatase crystal structure, while the IO-bPEI corresponded to a cubic spinel crystal structure. The designed nanocomposite, confirmed by XRD and SAED analyses, showed both crystal structures. The IO-bPEI NPs (individual NPs were about 13 nm in size) aggregates formed hetero-agglomerates with TiO₂@CA clusters (individual NPs were 3–5 nm in size), resulting in a composite size ranging from hundreds of nanometers to micrometers. SEM and elemental mapping confirmed the formation and elemental distribution of the elements (i.e., Ti, Fe, O, N, C) in the nanocomposite. Infrared spectroscopy identified typical functional groups, while XPS provided a comprehensive understanding of the elemental states in IO-TiO₂. Zeta potential measurements revealed a positive charge over a wide pH range for IO-TiO₂, primarily caused by IO-bPEI, which influenced the ciprofloxacin degradation process. Hydrodynamic diameter measurements, supported by TEM results, showed a significantly larger size of the nanocomposite (~2260 nm) compared to individual NPs. TGA confirmed the presence of organic layers on both the individual NPs and the IO-TiO₂ nanocomposite. Magnetic property analysis revealed a strong magnetic response with saturated magnification of 49 emu/g. BET analysis indicated a large specific surface area, which is crucial for enhanced photocatalytic processes and the formation of more reactive species. UV-Vis-DR measurements demonstrated the ability of the nanocomposite to generate charge carriers in the UV-B range, and PL results indicated reduced electron-hole recombination in the IO-TiO₂ heterojunction, thereby improving photocatalytic efficiency compared to TiO₂@CA alone, with twice the amount with respect to the latter in IO-TiO₂. To determine the optimal parameters for enhanced photocatalytic degradation of ciprofloxacin by IO-TiO₂, various effects were studied in detail, including the amount of nanocomposite, pH, time, and reusability. Optimal parameters for enhanced photocatalytic degradation of ciprofloxacin by IO-TiO₂ were determined: 10 mg IO-TiO₂, pH 7, and 2 h UV-B irradiation. The degradation process monitored by UV-VIS spectroscopy achieved up to 70% degradation of ciprofloxacin at a concentration of 8 mg/L after 150 min. Compared to the current state-of-the-art, the IO-TiO₂ nanocomposite exhibited acceptable and satisfactory performance with remarkable magnetic response and good reusability, maintaining stability and photocatalytic behavior over three cycles. The study shows that iron oxide not only acts as a carrier for TiO₂@CA NPs but also delays electron-hole recombination, which increases the photocatalytic efficiency. Overall, the successful preparation and performance of IO-TiO₂ nanocomposites highlight their significant potential as magnetic photocatalytic materials for mitigating antibiotic residues in wastewater, representing a key advance in combating antibiotics in environmental systems. Despite recent advances in the use of heterogeneous TiO₂-based materials for wastewater treatment, there are still technical issues that hinder its commercialization. Future studies will focus on the photocatalytic degradation of ciprofloxacin in the visible range with magnetically-based NPs and TiO₂ and on the industrialization of this technology to move from laboratory scale to real applications.