Synthesis of modified beta bismuth oxide by titanium oxide and highly efficient solar photocatalytic properties on hydroxychloroquine degradation and pathways.

With the outbreak of coronavirus pandemic the use of Hydroxychloroquine increased. These compounds have harmful effects on the environment, such as generation of antibiotic-resistant bacteria; therefore, their degradation has been considered as one of the environmental challenges. The purpose of this research is to synthesize heterogeneous structure of TiO2/β-Bi2O3 by hydrothermal method for solar degradation of Hydroxychloroquine. Then, the accurate characteristics of the synthesized samples were investigated by XRD, FESEM, TEM, XPS, UV-vis (DRS), and BET surface analyzer. Photocatalytic degradation of Hydroxychloroquine was studied under sunlight, and it was found that the visible light absorption of TiO2 photocatalyst by mixing β-Bi2O3 nanoparticles was greatly increased and 91.89% of the degradation was obtained in 120 min of photocatalytic reaction. This improvement can be attributed to the increased specific surface area, efficient charge transfer, and reduced electron-hole recombination with the β-Bi2O3 compound. Kinetic studies also reacted to follow pseudo-first-order kinetics. Also, demonstrated high stability and recyclability for nanoparticles, so that after 6 cycles of using the catalyst in take, 70.59% degradation was performed. According to the results, the excellent photocatalytic degradation activity demonstrated by the TiO2/β-Bi2O3, therefore, it is a potential candidate for the process of removing other organic contaminants from aqueous solutions.

Introduction

Water pollution is an important issue worldwide, and solving this problem requires the use of new principles to remove pollutants. The human need for water along with the limitation of water resources in terms of quantity and quality has led to the importance of purification and recovery of this vital substance [1]. Coronavirus (COVID-19) has become a global disaster and there is a strong search for effective drug treatment for it. In March 2020, hydroxychloroquine (HCQ) was recommended for the treatment of the new Covid-19 detected for the first time in China as of November 2019. In fact, in vitro studies have indicated a sure efficacy of HCQ against infection by COVID-19, and are now one of the most effective drugs in the treatment of COVID-19. Long-term use of hydroxychloroquine (HCQ) is a cornerstone in the treatment of this disease [2]. However, the entry of substances containing these pharmaceutical compounds into the environment occurs through the wastewater of the pharmaceutical industry, hospital effluents, and effluents from wastewater treatment plants, laboratory activities, and human and animal wastes [3]. The use of HCQ is causing unanticipated harm to the environment. When HCQ enters the environment, they will lead to pharmaceutical stability of pathogenic bacteria and cause extremely serious harm to aquatic organisms in the environment and human health. Due to the increasing spread of these drugs and the further entry of these compounds into the effluent and sewage, more than ever, the removal of drugs from water and wastewater sources needs attention. Pharmaceutical compounds are present in surface and groundwater, sewage and even drinking water in the µg/l and ng/l, and given that the health care and physical health of communities is an important and fundamental principle around the world, the degradation, and removal of these Materials are of particular importance [4]. Various technologies such as activated carbon adsorption, reverse osmosis, filtration, and biological methods are used to remove drugs. However, the results indicated that the use of the above methods causes the transfer of contaminants from one phase to another, their concentration, and finally the production of a new contaminant that requires further treatment [5]. Advanced oxidation (AOPs) is one of the most popular processes that is widely accepted today due to the low-cost and simplicity of the process to remove resistant organic pollutants [6]. Compared with other methods, the chemical oxidation technique chiefly attains the purpose of degrading contaminates via chemical reactions, which can completely degrade contaminants without generating pharmaceutical-resistant bacteria [7]. Advanced oxidation is a process based on the production of free radicals, especially hydroxyl, which successfully attacks contaminant molecules [8]. Hydroxyl radicals are produced in aqueous media using H2O2, UV/H2O2, UV/TiO2, UV/ZnO, and other methods [9]. In recent decades, semiconductor photocatalysis is due to its environmental friendliness, cost-effectiveness, high efficiency, and excellent stability, which can convert abundant and clean solar energy into chemical energy, so this process is considered as a green technology for Water and wastewater treatment [10], [11]. To date, numerous semiconductor photocatalysts such as TiO2, ZnO, C3N4, WO3, Ag3PO4, and SrTiO3 have been investigated. However, TiO2 is still considered a suitable photocatalyst due to its high chemical stability, low- cost, and environmentally friendly [12], [13]. Due to its wide band gap energy and its low activity in visible light, the use of this nanoparticle is limited. Researchers have proposed many strategies to increase the photocatalytic properties of visible light by increasing the use of solar energy and improving the efficiency of charge separation, including doping, loading of metal particles and other semiconductors [14]. The fabrication of a heterogeneous photocatalyst is one of the inventive approaches to overcome these limitations, because doped semiconductor oxide can not only efficiently separate and transmit optical electrons-holes, but also extend the amplitude of the optical response by the narrow-band gap semiconductor [15]. Bismuth oxide (Bi2O3) is one of the heterogeneous components, with excellent attributes and narrow band gap (2 to 3.96 eV), high refractive index, good optical conductivity, and non-toxic properties, and high photocatalytic activity [16], [17]. To these properties, Bi2O3 is used in applications such as gas sensing, fuel cell, and water purification [18].

Bi2O3 emerges five fuzzy forms: α- (monoclinic), β- (tetn class="Chemical">ragonal), γ- (body-centered cubic), δ- (face-centered cubic) and ε- (triclinic) [19]. Among these forms, α-Bi2O3 is known as the constant phase at low temperature and δ-Bi2O3 is the stable phase at high temperature (729 °C), the other three forms of Bi2O3 have a metastable phase. β-Bi2O3, as a metastable phase, has recently been reported to have excellent photocatalytic activity relative to α-Bi2O3 [20]. β-Bi2O3 has the highest absorption in the visible light region because it has the lowest band gap (2.4 eV) and shows better photocatalytic performance under visible light radiation. Due to the β-Bi2O3 is an intrinsic p‐type semiconductor with high mobility, it can therefore be used as an electron donor in photocatalytic processes. TiO2 is an inherent n-type semiconductor, which can be used as an electron receptor and provides a pathway for the surface charge transfer. Based on this strategy in this study, a p-n heterogeneous photocatalyst is synthesized by binding β-Bi2O3 and TiO2. Whereas HCQ is the most widely used drug for the treatment of COVID-19 [2]. So, the substances containing these pharmaceutical compounds into the environment occurs [3]. When HCQ enters the environment, they will lead to pharmaceutical stability of pathogenic bacteria and cause extremely serious harm to the environment and human health. The aim of this study was to clarify the mechanism of the degradation of HCQ in aqueous solution and improve its photodegradation efficiency. Therefore, TiO2/β-Bi2O3 nanocomposite was synthesized by hydrothermal method and was used for the photodegradation of HCQ under solar light irradiation. Furthermore, the photodegradation mechanism of TiO2/β-Bi2O3 for HCQ was explored.

Materials and methods

Materials

Sodium hydroxide (n class="Chemical">NaOH), hydrochloric acid (HCl), bismuth nitrate (Bi(NO3)3·5H2O) and nitric acid (HNO3), citric acid (C₆H₈O₇), ethanol and titanium (IV) areopropoxides purchased from Merck, and was used without further purification. Hydroxychloroquine sulfate (HCQ) was prepared from Mofid Pharmaceutical Company. It should be noted that deionized water was used during photocatalytic synthesis and degradation experiments.

Preparation of β-Bi2O3 nanoparticles

Initially for the synthesis of β-n class="Chemical">Bi2O3, 00.9 g Bi (NO3)3·5H2O was dissolved in 20 ml of HNO3 at a concentration of 1 M. Then 0.3 g of citric acid was added to the solution and after stirring for 1 h, the NaOH solution was added dropwise until the pH of the solution reached 4. The resulting suspension was then placed at room temperature for 1 h. The suspension was then placed in an autoclave and then placed for 2 h at 180 °C. After the reaction, the resulting suspension placed at room temperature for 90 min to cool. The production precursor was centrifuged at 2500 rpm for 15 min and washed three times with deionized water and ethanol and then dried for 60 h at 60 °C. Finally, the precursor was calcined for 3 h at 350 °C and finally, β-Bi2O3 nanopowders were obtained [21].

Preparation of TiO2/β-Bi2O3 nanocomposite

The preparation process of TiO2/β-Bi2O3 nanocomposite was performed by hydrothermal method. For this purpose, in the first step, 10 ml of ethanol along with 20 ml of distilled water was added dropwise to 5 ml of titanium (IV) isopropoxide solution and stirred for 30 min. Then 5 ml of acetic acid was added to the resulting solution and finally 40% Bi2O3 was added to TiO2. The resulting solution was placed in an autoclave at 160 °C for 24 h. Finally, after centrifugation at 2500 rpm for 20 min, the solution was washed several times with distilled water and placed in an oven at 80 °C for 8 h [22].

Investigation of physical properties of synthesized catalysts

The crystal structure of the TiO2/β-n class="Chemical">Bi2O3 nanocomposite synthesized by the Rigaku MiniFlex 600 (XRD) radiation diffract meter was investigated using Cu Ka radiation as the X-ray source. Fourier transform infrared (FTIR) spectroscopy (Shimadzu, FTIR 1650 spectrophotometer, Japan) were also applied to determine the chemical features of the nanocomposite. X-ray photoelectron (XPS) spectroscopy was performed using AMICUS, Kratos Analytical (Shimadzu) spectroscopy to investigate the oxidation state and chemical environment of the elements in the sample. The morphology and morphology of the samples were studied by a field diffusion scanning electron microscope (FESEM; JEOLJSM-7600F) equipped with an energy scattering spectrometer (EDS). The detailed study of the structure and size of nanoparticles was performed using TEM analysis (Philips EM208S 100KV). Nanocomposite topography was determined using atomic force microscopy (AFM). Ultraviolet reflection spectrum (DRS) was recorded on UV–Vis spectroscopy (Shimadzu, UV-2550, Japan). The Brunauer-Emmett-Teller (BET) surface area and sample pore size distribution were studied in the N2 adsorption analyzer (NOVA 2000e) the USA. Active free radicals were determined using electron paramagnetic resonance (EPR) in a Bruker ELEXSYS 500 spectrometer.

Photodegradation tests

A batch system was used to perform photodegradan class="Chemical">tion experiments. The light used in the experiment was solar light and UV-A light. The batch photocatalytic system in this study was exposed to direct sunlight with an intensity of 76–72 kW and UV-A light with a light density of 6 W. The parameters studied in this experiment include pH different [3], [5], [7], concentration of TiO2/β-Bi2O3 nanocomposite as a catalyst (0.05, 0.1, 0.3, 0.4 g/L), hydroxychloroquine concentration (1, 10, 20, 40 mg/L), H2O2 concentration (1,2, 3, 4 mg.L−1), temperature (10, 20, 40, and 60 °C) and irradiation time (5,10,15,20,30,45,60,90,120 min). (0.1 M) HCl and (0.1 M) NaOH were used to adjust the pH, and a magnetic stirrer at 70 rpm was used to blend the system. Photodegradation experiment in dark conditions and photolysis was also performed. It should be noted that before exposing the sample to sunlight, and UV-A light irritation, the solution was first stirred for 30 min in the dark to ensure an adsorption-adsorption balance between the TiO2/β-Bi2O3 photocatalyst and the hydroxychloroquine drug. Over a period, 5 ml of the solution was extracted. After centrifugation at 2000 rpm for 15 min, the adsorption rate of hydroxychloroquine was measured by a spectrophotometer (UV-770 Brite) at 343 nm. Equation [1] was used to calculate the HCQ removal percentage [23]:

In the above relation, n class="Chemical">C0 is the initial concentration of hydroxychloroquine sulfate, Ce is the final concentration of hydroxychloroquine sulfate.

Results and discussion

Structural characterizations of TiO2/β-Bi2O3

XRD analysis

The crystalline properties of β-n class="Chemical">Bi2O3, TiO2, and TiO2/β-Bi2O3 nanoparticles were analyzed using the XRD method. As Fig. 1 shows all diffraction peaks at 19.55,21.76, 25.63, 27.4, 32.15, 35.86, 37.35, 46.85, 48.46, 54.78 and 63.45°belong to the surface of (1 1 1), (0 2 0), (0 0 2), (1 2 1), (2 1 2), (1 1 3), (0 4 1), (1 0 4), (2 4 1), (2 2 5), and (1 6 1) crystals, which corresponds to the tetrahedral β-Bi2O3 (JCPDS No. 71–2274) [24]. The XRD peaks in TiO2 show anatase peaks (JCPDS 00–021-1272) so that diffraction peaks are located at 25.63, 32.15, 35.85, 47.5, 53.85, 54.83, 57.15, 62.43°, which correspond to (1 0 1), (2 2 2), (3 1 1), (2 0 0), (4 2 2), (1 0 5), (5 1 1) and (2 0 4) anatase TiO2 [10]. Peaks of β-Bi2O3, as well as TiO2, were observed in the TiO2/β-Bi2O3 nanocomposite.

Fig. 1

XRD patterns for β-Bi2O3, TiO2, and TiO2/β-Bi2O3 nanocomposite.

XRD patterns for β-Bi2O3, n class="Chemical">TiO2, and TiO2/β-Bi2O3 nanocomposite.

Fourier transform infrared (FTIR)

Fig. 2 shows the absorption peaks of the β-n class="Chemical">Bi2O3, TiO2 and TiO2/β-Bi2O3 samples under the FTIR. The peak at 530 (cm−1) corresponds to the symmetric tensile states of the Bi-O bond of Bi2O3 species [25]. In addition, a significant decrease is observed in the FTIR spectrum for the vibration mode at 1200–1700 (cm−1), which is a characteristic of the NO3 group [26].The absorption peak at 1086 (cm−1) was attributed to group C-O. The absorption peak in the range of 1384 (cm−1) can be observed to Ti-O vibration in TiO2 [27]. Also, the peaks in 1618, 28 and 3252 (cm−1) are attributed to O–H vibrations in the hydroxyl group that is absorbed by the water at the surface [28].

Fig. 2

FTIR spectra for β-Bi2O3, TiO2, and TiO2/β-Bi2O3 nanocomposite.

FTIR spectn class="Chemical">ra for β-Bi2O3, TiO2, and TiO2/β-Bi2O3 nanocomposite.

The presence of NO3 in the Fn class="Chemical">TIR spectrum is due to the thermal analysis of Bi (NO3)3 in the initial composition of bismuth oxide. Therefore, the absorption bands at 1200–1700 (cm−1) show the presence of NO3 groups [26]. Therefore, it can be stated that the hydroxyl and carboxyl, and NO3 groups, respectively, are located in larger amounts on the surface of the particles of the samples, and according to the resulting nanoparticles it can cause better stability in the aquatic solution.

FESEM –EDX analysis for TiO2, β-Bi2O3, and TiO2/β-Bi2O3

The TiO2, β-n class="Chemical">Bi2O3, and TiO2/β-Bi2O3 nanoparticles are porous accumulates composed of smaller crystallizes. TiO2 crystals have an average size of about 46 nm nanometers with a spherical shape (Fig. 3 ). β-Bi2O3 is formed as nanoparticles with an average size of about 37 nm (Fig. 3). FESEM images distribute the presence of spherical β-Bi2O3 nanoparticles evenly on the surface of TiO2 particles. As can be seen, the particles using this preparation method are somewhat accumulated, which this accumulates may be attributed to the evaporation of organic matter adsorbed on the nanoparticle surface and structural changes as well as the lack of sonicates of the samples before imaging [5]. It can be concluded that, under the current path, bismuth metal nanoparticles are formed by hydrothermal reaction and after the calcination is completely transformed into TiO2/β-Bi2O3 composite nanoparticles. In Fig. 3, the EDX spectra of the synthesized TiO2/β-Bi2O3 compounds showed that Bi2O3 is well distributed in TiO2 and there appears to be no separation of the two oxide phases. As can be seen, EDX analysis confirms the presence of β-Bi2O3 nanoparticles in TiO2.

Fig. 3

FESEM images of TiO2 nanoparticles, β-Bi2O3, TiO2/β-Bi2O3 nanocomposite and EDX analysis of TiO2/β-Bi2O3 nanocomposite.

FESEM images of TiO2 nanoparn class="Chemical">ticles, β-Bi2O3, TiO2/β-Bi2O3 nanocomposite and EDX analysis of TiO2/β-Bi2O3 nanocomposite.

TEM analysis for TiO2, β-Bi2O3, and TiO2/β-Bi2O3

TEM images of composite nanoparticles (Fig. 4 ) showed that n class="Chemical">TiO2 nanoparticles had a spherical shape with a diameter of approximately 67 nm. Fig. (4b), darker crystallites calculating about 42 nm in size are formed in the size of β-Bi2O3 crystals. Also, β-Bi2O3 nanoparticles were unstable under the electron rays of TEM and underwent structural arrangement. TEM images (4c) confirmed the presence of spherical β-Bi2O3 nanoparticles 16 nm in diameter uniformly on the surface of TiO2 particles.

Fig. 4

TEM image of (a) TiO2; (b) -β-Bi2O3; (c) TiO2/β-Bi2O3 nanoparticles.

TEM image of (a) TiO2; (b) -β-n class="Chemical">Bi2O3; (c) TiO2/β-Bi2O3 nanoparticles.

Atomic force microscopy (AFM) analysis for surface analysis of TiO2/β-Bi2O3 layers

AFM analysis is possible to determine the roughness and predicted surface parameters of n class="Chemical">TiO2/β-Bi2O3 layers. Fig. 5 shows an AFM micrograph of the surface of TiO2/β-Bi2O3 layers. As can be seen, the surface of the TiO2/β-Bi2O3 layers with Ra = 152 nm and Rq = 191 nm show a calculated surface area of 1.318 μm. This roughness is essential due to the dendritic growth of TiO2/β-Bi2O3 layers to increase HCQ uptake and photocatalytic mineralization [29].

Fig. 5

AFM analysis of topographic micrographs of TiO2/β-Bi2O3 surface layers.

AFM analysis of topographic microgn class="Chemical">raphs of TiO2/β-Bi2O3 surface layers.

XPS analysis for TiO2/β-Bi2O3 nanoparticles

Existing compounds and chemical states of n class="Chemical">TiO2/β-Bi2O3 composite nanoparticles were investigated using XPS analysis. As can be observed, the spectrum confirmed the presence of C, O, Ti, and Bi at the surface of β-Bi2O3/TiO2. Fig. 6(a) (a) shows the XPS spectrum of TiO2/β-Bi2O3. The binding energy spectrum of Ti 2p at 458.2 and 1/464 eV corresponds to Ti 2p3/2 and Ti 2p1/2 (Fig. 6 b) [5]. Bi 4f symmetric peaks with axes of 164.1 and 159.5 eV are assigned to β-Bi2O3 (Fig. 6 c). Peaks at 539.5 and 530.4 eV were indicative of oxygen in β-Bi2O3 and TiO2 (Fig. 6 d) [24].

Fig. 6

XPS spectrum TiO2/β-Bi2O3 nanocomposite. (a) Complete investigation of TiO2/β-Bi2O3, (b) Ti, (c), Bi and (d). Oxygen.

XPS spectrum TiO2/β-n class="Chemical">Bi2O3 nanocomposite. (a) Complete investigation of TiO2/β-Bi2O3, (b) Ti, (c), Bi and (d). Oxygen.

Brunauer – Emmett – Teller (BET) level analysis for TiO2/β-Bi2O3 nanocomposite

One of the effective factors in photocatalyn class="Chemical">tic performance is BET sample specific regions, which are measured by nitrogen adsorption–desorption method. The samples synthesized in this study, according to the International Union of Pure and Applied Chemistry (IUPAC) classification, have mesoporous surface and show a type IV isotherm with a type H3 residue. The BET level for the composite sample was 156.7 m2/g. It was also observed from the pore size distribution of BJH that the resulting sample has a narrow pore size distribution. The total pore volume was 0.394 cc/g. In this study, the average pore radius was 13.76 nm Fig 7 .

Fig. 7

Adsorption-desorption isotherm N2 for TiO2/β-Bi2O3. And BJH pore size distribution.

UV–vis spectrum analysis (DRS) for TiO2, β-Bi2O3 and TiO2/β-Bi2O3 nanocomposite

The UV–vis spectrum (DRS) was obtained from TiO2, β-n class="Chemical">Bi2O3 and TiO2/β-Bi2O3 samples and the results are shown in Fig. 8 . Since TiO2 showed high absorption in the UV region (200–400 nm) with a band gap of 3.2 eV, which is typical for TiO2 anatase. While β-Bi2O3 had more absorption in the visible light region, which showed a band gap energy of 2.46 eV [30]. While the UV spectrum of TiO2/β-Bi2O3 composites is absorbed more light in the visible light region compared to TiO2 and β-Bi2O3 [31]. Therefore, the highest visible light response among nanoparticles was related to TiO2/β-Bi2O3 nanocomposite, which is attributed to the synergy effect between TiO2 and β-Bi2O3. Therefore, this nanocomposite can be used in photocatalytic reactions due to better performance and more efficient use of light [12].

Fig. 8

The UV–vis spectrum (DRS) spectrum for the synthesized samples.

Analysis of electron paramagnetic resonance (EPR)

The recognition •OH in aqueous media was investigated by electron paramagnetic resonance (EPR) without adding pollutants. To investigate the role of active oxygen free radicals in the photocatalytic degradation of HCQ, trapping experiments with isopropyl alcohol (IPA), methyl alcohol (MA), and ammonium oxalate (AO) were added to trap the ·O2 −, ·OH, and h+. As Fig. 9a shown, in the absence of sunlight irradiation, there was no definite peak relating to DMPO− • OH, and in the ERP spectra, demonstrated that no ·OH will be produced in the absence of light irradiation of TiO2/β-Bi2O3. After sunlight irradiation, the EPR spectrum with an intensity ratio of 1:2:2:1 signal was clearly observed, which related to the DMPO−•OH [10]. Meanwhile, •O2 − was detected in methanol solution; the result is indicated in Fig. 9 (b). Without sunlight irradiation, there was no signal connected by DMPO- •O2 − appeared in the EPR spectrum. While, by irradiating sunlight to the TiO2/β-Bi2O3 composite, a four-line special signal with an intensity ratio of 1: 1: 1: 1 is created in the EPR spectrum, which relates to the peak of DMPO- • O2 −. These results showed that •OH and •O2 − were produced in TiO2/β-Bi2O3 in the photocatalytic process with sunlight irradiation.

Fig. 9

EPR spectra of TiO2/β-Bi2O3 a) DMPO-·OH and b) DMPO-·O2−.

EPR spectra of n class="Chemical">TiO2/β-Bi2O3 a) DMPO-·OH and b) DMPO-·O2−.

Solar light–induced photodecomposition of HCQ

Effect of pH on photocatalytic degradation of HCQ

The pH of the reaction medium plays an important role in the oxidation process by the photocatalytic reaction. The pH value has a decisive effect on the oxidation potential of OH radicals due to the correlation between the oxidation potential and the pH value. Therefore, the role of pH in any photocatalytic reaction must be determined [19]. Therefore, in this study, the impact of pH on the degradation of HCQ using TiO2/β-Bi2O3 catalyst in the pH range of 3–11 with a dose of 10 mg/L HCQ, 0.1 g/L catalyst and 0.1 mg.L−l H2O2 was examined at ambient temperature and in direct solar light irradiation. Fig. 10 shows the effect of pH on the degradation of hydroxychloroquine in the pH range from 3 to 11. The highest drug degradation was obtained by using a catalyst at pH 3 with 92.37%. It is assumed that the surface of β-Bi2O3/TiO2 catalyst is positively charged in acidic solution while negatively charged in alkaline solution [32]. Also, the drug hydroxychloroquine is known as an organic substance that has a negative charge. Therefore, it is clear why the highest degradation is obtained at acidic pH compared to alkaline pH. The activity of the catalysts may be due to the presence of a strong electrostatic field between the surface of the positively charged catalyst and the negatively charged drug. This finding showed the important role of pH in the degradation of organic pollutants. Another reason for the increase in degradation in an acidic medium is the production of hydroxyl radicals, which were higher than in an alkaline medium, so these radicals can increase the oxidation potential. In addition, the oxidation potential due to the recombination of hydroxyl radicals in an acidic medium is lower than that of an alkaline [1]. At low pH, the adsorption of cationic organic molecules on the photocatalyst surface increases, because the photocatalyst surface has a positive charge, which leads to increased adsorption of cationic organic molecules. Therefore, the degradation efficiency increases at low pH .

Fig. 10

Effect of pH on HCQ degradation.

Effect of pH on HCQ degn class="Chemical">radation.

Effect of temperature on photodegn class="Chemical">radation of HCQ.

Effect of temperature

The effect of temperature on the optical degradation of hydroxychloroquine molecules was investigated using TiO2/β-Bi2O3 nanoparticles Fig 11. As the temperature increases, the photocatalytic degradation efficiency of HCQ molecules increases. Therefore, increasing the temperature has a natural effect on chemical reactions and increases the degradation process in general by changing the activation energy. An increase in temperature led to an increase in molecules. Kinetic energy had a positive effect because it helps them to move molecules to active regions. As the temperature rises, bubbles formed in the solution, leading to the production of free radicals. In addition, increasing the temperature helps the degradation reaction to overcome a combination of electron holes. Also, increasing the temperature may increase the rate of oxidation of the interface organic molecules [33]. However, with further increases in temperature, the more kinetic energy produced that is able to move the polluting molecules away from the active regions before reaching the degradation process. Therefore, the optimum temperature is 40 °C and after this temperature, the inverse relationship between optical degradation and temperature occurs due to the increase in kinetic energy of the molecules. As a result, the rate of degradation decreased with decreasing hydroxyl production in the solution.

Fig. 11

Effect of temperature on photodegradation of HCQ.

Effect of hydrogen peroxide (H2O2)

By adding some irreversible electron receptors, including H2O2, to the reaction mixture, the recombination of electrons and cavities in photocatalytic reactions can be reduced [34]. Thus, in order to evaluate the ability of photocatalytic degradation with the assistance of H2O2, an experiment was performed under catalytic dose of 0.1 g/L, 10 mg/L of the HCQ, and pH 3 with 1–4 mg. L−1 H2O2under sunlight. The results (Fig. 12 ) showed that sunlight, catalyst, and H2O2 together have a significant effect on the degradation process. Increasing the concentration of oxygen in the solution was very useful for the photocatalytic degradation of the drug. As can be seen, with increasing H2O2 concentration, the degradation of pollutant molecules accelerates·H2O2 traps electrons and thus prevents the recombination of electron-hole pairs, thereby increasing H2O2 probability of OH and O2 radical formation at the photocatalyst surface. Beyond the optimal concentration of H2O2, the rate of hydroxychloroquine degradation reduced due to the radical reduction of OH [1], [35]. Subana and Swaminathan reported that increasing the amount of H2O2 beyond the optimal level produces more peroxide radicals, which acts as a scavenger for the hole and thus reduces the efficiency of photocatalytic degradation [34]. The accepted mechanism for H2O2 is to fracture the O-O bond with the function of sunlight, which causes the formation of a hydroxyl radical (•OH), which leads to drug degradation [1].

Fig. 12

Effect of different concentrations of H2O2 on the photodegradation of HCQ.

The effect of H2O2 dose on n class="Chemical">HCQ degradation can be explained in the number of OH radicals produced and to the absorbance of OH radicals [36]. It is well known that H2O2 can trap photoelectron (e−) to electron-hole pairs (e− - h+). Additional OH radicals can be generated by the reaction between H2O2 and e− or •O2 − (Equations [3], [4]. Thus, the addition of H2O2 to the photocatalytic system leads to the degradation of HCQ. However, beyond the optimal dose, H2O2 traps OH radicals to produce weaker oxidizing radicals. Accordingly, trapping of OH radicals occurred through ((Equations. [5], [6]. The decrease in the concentration of radicals •OH produced, due to the higher dose of H2O2, inhibited the HCQ degradation. Likewise, the addition of H2O2 appears to act as a source of oxygen.

Effect of different doses of catalyst on photodegradation of HCQ

The photocatalytic activities of the catalysts under direct sunlight were also evaluated and compared. The optical decomposition curves of HCQ in water as a function of irradiation time in the absence and presence of different doses of catalyst are shown in Fig. 13. Before starting the photocatalytic tests, the solution was kept in the dark for 30 min. Photolysis trials in the absence of catalyst were also performed. As can be seen, the amount of degradation under photolysis and adsorption tests were slight (Fig. 13 a, and b). The solution was then subjected to photocatalytic tests and direct sunlight to evaluate the effect of different doses of catalyst. (Fig. 13c). The photodegradation efficiency achieved at different photocatalyst concentrations of 0.05, 0.1 and 0.2, 0.3 and 0.4 g/L, respectively (38.15), (52.36), (63.57), (91.45) and (80.42). As can be observed, the amount of degradation increased with the increase of the catalyst. Due to the increase in active sites, the rate of photocatalytic degradation of HCQ increases with the photocatalytic dose. The photocatalytic removal mechanism based on the production of OH radicals. Thus, the HCQ contaminant removed through these radicals [37], [38]. But beyond the amount (0.3 g/L), due to the increase in turbidity in the solution, the amount of degradation decreased, and thus due to the increase in light scattering and the lower penetration depth of photons, fewer photocatalysts can be activated. In addition, the accumulation of nanoparticles at high concentrations leads to a reduction in the number of surfactant sites available for photocatalytic degradation [39] and inactivation of activated molecules, resulting in the collision of activated molecules with ground-state molecules [40].

Fig. 13

a: Photolysis, b: Adsorption, and c: Effect of photocatalyst concentration on photocdegradation of HCQ.

a: Photolysis, b: Adsorption, and c: Effect of photocatalyst concentn class="Chemical">ration on photocdegradation of HCQ.

Effect of initial concentration HCQ photocatalytic degradation of HCQ

The initial concentn class="Chemical">ration of hydroxychloroquine from 1 to 40 mg/L at pH 3, 0.3 g/L catalyst at a maximum illumination time of 120 min investigated, and the results shown in Fig. 14 . As can be seen, increasing the initial concentration of HCQ reduced the degradation efficiency in photolysis and photocatalytic processes. Increasing the HCQ from 1 mg/L to 40 mg/L reduced the removal efficiency of HCQ from (91.89) at a dose of 10 mg/L to (44.26 mg/L) using a catalyst of 0.3 g/L. The reason for this decrease is that increasing the HCQ concentration prevents the efficiency of HCQ removal in the photocatalytic process. This is due to the probability of HCQ contact at a lower concentration is relatively high. However, the relative availability of surface-active sites decreases with increasing HCQ concentration [41]. Higher concentrations of HCQ increased the inhibitory effect and thus reduced the degradation efficiency [38]. Similar results for the photocatalytic degradation of other dye and drug contaminants showed that the surface of catalysts saturated at higher concentrations of contaminants [1].

Fig. 14

Effect of initial HCQ concentration on photodegradation of HCQ.

Effect of initial n class="Chemical">HCQ concentration on photodegradation of HCQ.

HCQ degradation kinetics

The time-dependent degradation of HCQ studied under a photocatalytic process. This provides the degradation efficiency and constant stability of the degradation processes. The Kinetics were studied as a function of soluble HCQ concentration. In This study, the optimal experimental conditions including pH = 3, 10 mg/L concentration of HCQ, catalyst concentration of 0.3 g/L and time of 120 min were considered. C/C0 values were plotted as a function of time for all concentrations of HCQ under optimal conditions. The results are shown in Fig. 15 . The figure clearly shows that an initial sharp decrease in Ct/C0 in the photocatalytic degradation of HCQ occurred with increasing concentration. The kinetics of the HCQ photodegradation reaction in the presence of the TiO2/β-Bi2O3 catalyst were calculated via Eq.where C is the instantaneous concentration and C0 is the initial C, K is the constant reaction rate and t is the reaction time [42]. As Fig. 15 shows, the photodegradation of HCQ is linear over time and followed pseudo-first-order kinetics. Over time, the rate of HCQ degradation reaction decreased from (0.0414) to (0.0129) min−1 (Table 1 ).

Fig. 15

HCQ degradation kinetics.

Table 1

Parameters affecting HCQ degradation kinetics.

Equation	K(min−1)	R2	HCQ(mg/l)
Y = 0.0414x + 1.7822	0.0414	0.983	1
Y = 0.0334x + 1.4182	0.0334	0.988	10
Y = 0.0221x + 1.1221	0.0221	0.965	20
Y = 0.0129x + 0.3524	0.0129	0.973	40

HCQ degn class="Chemical">radation kinetics.

Parameters affecn class="Chemical">ting HCQ degradation kinetics.

Comparison of photocatalytic degradation of HCQ under UV and sunlight

Due to the difference in photon energy in UV and sunlight, the photocatalytic activity of catalysts may be affected by UV sources. Hence, UV-A was used as a 6-watt UV source for comparison with sunlight in the TiO2/β-Bi2O3 photocatalytic system. The conditions are as follows: 10 mg/L HCQ, 0.3 g/L photocatalyst, pH 3 and 40 °C for 120 min and the results are presented in Fig. 16 . The results shown that HCQ degradation occurred with higher efficiency in sunlight compared to UV-A light. So that 90.27% of the degradation was observed in 120 min under UV-A light, while in the presence of sunlight for 120 min with a radiation intensity of 76–72 klux, the degradation was higher with UV-A and 91.89%. In fact, higher intensities of sunlight determine the amount of photons absorbed by the catalyst. By increasing the duration of light irradiation, the catalyst absorbs more photons, which produces more electron-hole pairs on the surface of the catalyst, which increases the concentration of hydroxyl radicals, resulting in higher removal rates. Therefore, it can be inferred that solar photocatalytic degradation can reduce energy costs and stabilize them in areas with high sunlight. In this study, sunlight was more effective than UV light in degrading HCQ contaminants. Solar energy can be substituted as an effective light source due to its abundance, freeness, renewability and non-hazardous nature [35], [43].

Fig. 16

Comparison of sunlight and UV-A light in HCQ photodegradation.

Comparison of sunlight and UV-A light in HCQ photodegn class="Chemical">radation.

A plausible mechanism for reaction

The valence band edge potential (n class="Chemical">EVB) and the conduction band edge potential (ECB) determine the photocatalytic reaction. EVB was calculated by the following experimental formula. where EVB is the edge potential of EVB , Ebg is the energy band gap, Ee is the energy of the free electrons on the hydrogen scale (about 4.5 eV), and X is the electronegativity of the semiconductor. Ebg is the energy of the semiconductor band gap, which in Fig. 17 corresponds to the DRS spectrum for TiO2 and β-Bi2O3 of 3.2 and 2.46 eV, and the corresponding electronegativity are 5.81 and 6.22 eV. Which was used to calculate the edge of the conduction band (ECB) in the formula [7].As a result, Measured EVB and ECB for TiO2 were 2.91 and 0.29 eV, respectively, and 2.94 and 0.49 eV for β-Bi2O3. Since the conduction band edge for β-Bi2O3 nanoparticles was low (0.49 eV), it is not able to provide the sufficiently negative potential for optically excited electrons to reduce the absorbed O2 . In addition, the ECB for TiO2 was calculated higher than of the H+/H2 decline potential [15]. Therefore, the edge band potential (ECB) of TiO2 is more active than β-Bi2O3. This difference in the conduction band and valence band potentials leads to a potential difference between the two semiconductor catalysts, resulting in the formation of a heterojunction. As a result, the optically generated electron (e−) migrates from the TiO2 conduction band to the β-Bi2O3 conduction band. Due to the fact that the holes (h+) move in the opposite direction, the holes on the surface of β-Bi2O3 can be transferred to TiO2 via the potential difference. Therefore, this heterojunction created the interface of the two semiconductors leads to efficient charge separation, thus reducing the recombination of charge carriers (44). Also, on the other hand, the holes produced in VB of TiO2 can be transferred to VB β-Bi2O3 and lead to an increase in the longevity of the generated light charge carriers. In the catalyzed reaction of sunlight, there are many productions of electrons and holes and other oxidative species (ROS). Hence, the photocatalytic efficiency is increased and the HCQ is completely degraded. The schematic reaction pathway on the surface of the TiO2/β-Bi2O3 photocatalyst is shown in Fig. 17.

Fig. 17

Proposed mechanism for photocatalytic degradation of HCQ.

Proposed mechanism for photocatalytic degn class="Chemical">radation of HCQ.

Photocatalytic degradation of HCQ at large scale

In this study, the photodegradan class="Chemical">tion of HCQ under a larger-scale photocatalytic system was also investigated Fig 18. For this purpose, a 500 ml container containing HCQ contaminant solution, pH = 3, catalyst concentration of 0.3 g/L, HCQ dose of 10 mg/L was used. The experiment showed that the degradation efficiency of HCQ increased with the passage of time after 120 min, the HCQ degradation efficiency under solar light at large-scale was achieved 70.59%. The results are consistent with the study of Ahmadpour et al. [32] .

Fig. 18

Impact of large scale on HCQ degradation by TiO2/ β-Bi2O3.

Impact of large scale on HCQ degn class="Chemical">radation by TiO2/ β-Bi2O3.

Stability and reusability of the photocatalyst

The reusability of the n class="Chemical">TiO2/β-Bi2O3 photocatalyst and the removal efficiency of HCQ contaminant was determined as a promising process and a suitable strategy for photocatalytic degradation of the contaminant in each photocatalytic cycle. For this purpose, after each photocatalytic cycle, the TiO2/β-Bi2O3 photocatalyst was separated from the solution and washed several times with distilled water and ethanol, and then dried in an oven at 80 °C for 8 h, and the next step was to remove the HCQ was used [12]. As shown in Fig. 19 , it was noticed that upon reusing the nanoparticles for several cycles, the performance of nanoparticles remains the same and there is no remarkable loss in efficiency. The TiO2/β-Bi2O3 catalysts have a relatively stable and reusable activity, so that after 6 cycles of reuse of the photocatalyst, the photocatalytic degradation of HCQ increased from 91.89% in the first stage to 78.53% decreased in the sixth stage. According to the results, it can be stated that the TiO2/β-Bi2O3 nanoparticles are not disabled during photocatalytic reactions and its application is very economical due to its potential reuse [1].

Fig. 19

a) The reusability of TiO2/β-Bi2O3 nanoparticles, b) FTIR spectra of TiO2/β-Bi2O3 a: after and b: before 6 cycles photodegradation, and c) XRD pattern of the photocatalyst before and after 6 cycles photodegradation.

a) The reusability of n class="Chemical">TiO2/β-Bi2O3 nanoparticles, b) FTIR spectra of TiO2/β-Bi2O3 a: after and b: before 6 cycles photodegradation, and c) XRD pattern of the photocatalyst before and after 6 cycles photodegradation.

The superior activity of the prepared photocatalyst originates from the comn class="Chemical">binatorial factors viz., ordered mesoporous support with high specific surface area and optimum acidity originated from the TiO2/β-Bi2O3; the promotional impact of TiO2 and combination with β-Bi2O3 phase submitted the synergistic advantage. In other words, this high photocatalytic performance can be attributed to the synergistic effect between TiO2 and β-Bi2O3 nanoparticles, narrow bandgap force, and carrier separation [12]. The optimal pH is to be preserved during the photocatalyst synthesis is an important factor to tune the physicochemical characteristics of the photocatalyst and attain high activity.

As mentioned, by changing the pH to acidic the intrinsic and extrinsic attributes of the photocatalyst such as chemical adsorpn class="Chemical">tion, surface acidity, surface composition, reducibility, surface morphology, and surface coordination are significantly modified.

XRD and FTIR analysis were performed after 6 reuse cycles to invesn class="Chemical">tigate the crystal structure and change the chemical composition of TiO2/β-Bi2O3. The XRD results before and after degradation demonstrated no obvious change in the crystalline structure of TiO2/β-Bi2O3 after six recycles (Fig. 19c). Besides, the FTIR analysis of TiO2/β-Bi2O3 was performed before and after HCQ degradation (Fig. 19b). As Fig. 19b shows, there was not exhibited much change between the four FTIR patterns before and after degradation. According to Fig. 19b, the additional peaks of 3372 cm−1 related to O–H tensile vibrations, 1177 cm−1 pertained to C-O stretching vibrations, peaks of 1824 and 2767 cm−1 pertained to C–H stretching vibrations were observed in the 6th reuse. As noticed, much alteration did not occur in the FTIR peaks before and after the degradation process. This result indicated that TiO2/β-Bi2O3 had excellent chemical stability and can be used as a reusable photocatalyst in most applications.

Mineralization of HCQ

Photodegradan class="Chemical">tion of HCQ and mineralization with 0.05–0.4 g/l TiO2/β-Bi2O3 catalyst under optimum conditions was evaluated. Fig. 20 indicates the mineralization efficiency within 120 min. As can be seen, the mineralization efficiency (67.84) was acquired in 120 min. The increase in mineralization efficiency in the first hour of light irradiation is due to the absorption of light from the HCQ molecule or the formation of intermediates in the solution. Therefore, it can be seen that the intermediates produced in the solution are not sensitive to photolysis. This means that HCQ has been degraded by the photolysis process, but the sensitivity of intermediaries due to degradation to photolysis was lesser [10,4].

Fig. 20

Mineralization efficiency of HCQ using different TiO2/β-Bi2O3 under optimal conditions.

Mineralizan class="Chemical">tion efficiency of HCQ using different TiO2/β-Bi2O3 under optimal conditions.

Conclusion

In summary, visible light-responsive TiO2/β-n class="Chemical">Bi2O3 composite nanoparticles were successfully synthesized by the hydrothermal method and used in the degradation of hydroxychloroquine from aqueous solutions. The results showed that these synthesized composite nanoparticles possessed well crystallinity, significant optical properties, and high surface area. This nanocomposite has good nanostructures and dispersion, resulting in a remarkable photocatalytic performance in the photocatalytic degradation of hydroxychloroquine under direct sunlight and UV light. In addition, an acceptable mechanism was proposed for the photocatalytic degradation of hydroxychloroquine by TiO2/β-Bi2O3, and OH and O2 − radicals were important to the photocatalytic activity. This work not only presents a highly efficient photocatalyst for hydroxychloroquine degradation, but also provides a promising method to remove other contaminants from aqueous solutions for wastewater treatment.

Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relan class="Chemical">tionships that could have appeared to influence the work reported in this paper.