DOI: 10.31038/NAMS.2023622

Abstract

In this study, a novel polyaniline/graphitic carbon nitride/zinc tungstate (PANI/g-C₃N₄/ZnWO₄) (PGZ) ternary nanocomposites (NCs) as a heterostructure photocatalys was examined during photocatalytic degradation process in the efficient removal of glyphosate herbicide from a aqueous solution. Different pH values (3.0, 5.0, 7.0, 9.0 and 11.0), increasing glyphosate concentrations (5 mg/l, 10 mg/l, 15 mg/l and 20 mg/l), increasing PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations (5 mg/l, 15 mg/l, 30 mg/l and 45 mg/l) and increasing recycle times (1., 2., 3., 4., 5., 6. and 7.) was operated during photocatalytic degradation process in the efficient removal of glyphosate in a aqueous solution. The characteristics of the synthesized nanoparticles (NPs) were assessed using X-Ray Difraction (XRD), Field Emission Scanning Electron Microscopy (FESEM), Energy-Dispersive X-Ray (EDX), Fourier Transform Infrared Spectroscopy (FTIR), Transmission Electron Microscopy (TEM), Diffuse reflectance UV-Vis spectra (DRS) and X-Ray Photoelectron Spectroscopy (XPS) analyses, respectively. The cyctotoxicity test was operated to the standard TBE (trypan blue dye exclusion) assay technique with Drosophila melanogaster (fruit fly). ANOVA statistical analysis was used for all experimental samples. The maximum 99% glyphosate removal efficiency was obtained during photocatalytic degradation process in aqueous solution, at 15 mg/l gylphosate, at 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs, at pH=11.0, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time and at 25°C, respectively. The maximum 99% cyctotoxicity removal was observed at untreated glyphosate samples, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively. The maximum 99% cyctotoxicity removal was observed at 5 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst concentrations, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively. The study revealed the excellent minimization of cytotoxicity of glyphosate after photocatalytic degradation process with the PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst. As a result, the PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst is found to be non-cytotoxic irrespective of its quantity used. Finally, the combination of a simple, easy operation preparation process, excellent performance and cost effective, makes this a novel PANI/g-C₃N₄/CoMoO₄ ternary NCs heterostructure photocatalyst a promising option during photocatalytic degradation process in agricultural industry wastewater treatment.

Keywords

ANOVA statistical analysis, Cytotoxicity test, Diffuse reflectance UV-Vis spectra (DRS), Drosophila melanogaster (fruit fly), Electrochemical filtration process, Energy-dispersive X-ray (EDX), Field emission scanning electron microscopy (FESEM), Fourier transform infrared spectroscopy (FTIR), Gylphosate, Herbicites, Ionic liquid-assisted synthesis, Novel polyaniline/graphitic carbon nitride/zinc tungstate (PANI/g-C₃N₄/ZnWO₄) ternary nanocomposites, Pesticides, Photocatalytic degradation, Transmission electron microscopy (TEM), X-ray difraction (XRD), X-ray photoelectron spectroscopy (XPS)

Introductıon

The intense populational growth and industrial expansion in the most diverse segments of society have led to a substantial increase in the demand for drinking water supply and large-scale food production [1]. Thus, to increase productivity at an economically profitable level, the employment of agrochemicals has been widely used to combat pests and weeds [2]. With the enhanced use of a new variety of anthropogenic compounds towards industrialization, water pollution has increased substantially [3]. Anthropogenic compounds like synthetic pesticides and herbicides are often used in agricultural fields to protect crops. However, pesticides are characterized by low biodegradability, high bioaccumulative capacity arising from their physicochemical properties, and a long half-life, of 5-15 years, increasing their toxicity to the environment and humans [4,5]. Thus, pesticide persistence in soil, wastewater, ground, and surface water has proved to be a considerable environmental problem, and may be compounded along the food chain, reaching concentrations toxic to human health [6]. Due to their high stability, these compounds can contaminate areas distant from pulverization through water volatilization and soil absorption. Studies have associated exposure to compounds with hormonal changes in the immune, neurological and cardiac systems, as well as with the development of neoplasms [7,8].

For this purpose, diverse techniques, such as adsorption and advanced oxidative processes (AOPs), which include Fenton, photo-fenton, heterogeneous photocatalysis and ozonation systems, have been explored for removing and degrading biopersistent organic compounds [9-11]. AOPs are based on the generation of free radicals, e.g., hydroxyl (OH^●) and superoxide (O₂^{– ●}) radicals, which have high oxidizing power in an aqueous solution and are able to degrade pollutants into lower molecular weight intermediates and inorganic precursors [12,13]. Heterogeneous photocatalysis is an advanced oxidative process that occurs through the photoactivation (by sunlight or artificial light) of a semiconductor, which uses water molecules and dissolved oxygen as reagents of oxi-reduction reactions [14]. This technique is very efficient and promising for the degradation of organic pollutants, including dyes, drugs and pesticides [15,16]. Among the materials used, metallic nanooxides (zinc oxide, ZnO and titanium dioxide, TiO₂) have been largely employed due to their excellent properties, such as low toxicity, good availability, chemical stability, large surface area/porosity, and photocorrosion [17,18]. However, these conventional nanocatalysts are characterized by their high bandgap energy, which is the energy required to start photocatalytic reactions. Additionally, due to their high surface energy, they tend to agglomerate during the photocatalytic process. Therefore, the association of these nanocatalysts with a second, less active material (called catalytic support or matrix) can solve these drawbacks, even when the active material is dispersed in low concentrations (ca. 0.5-5 wt%) on the support [19]. Thus, combining the two materials results in a new material called NCs, in which the active substance is in the above-mentioned concentration range and this is named the reinforcement phase [20].

NCs are multiphase materials formed by a continuous and dispersed phase and have at least one dimension in the nanoscale [21]. The continuous phase (matrix) consists of a compound of polymeric, ceramic or metallic origin, while the dispersed phase (reinforcement) is commonly derived from fibrous materials [22-24]. NCs materials are synthesized to combine individual properties and reduce limitations, such as physicochemical and thermal instability, expanding the scope of applications. In parallel, at the nanoscale, the materials exhibit distinct behaviors to those found at the micrometer scale, such as volume/area relationship and increased reactivity [25]. Another technique widely used for pesticide removal from wastewater consists of adsorption, especially when using nanomaterials (adsorbents), due to its simplicity of operation, relatively low cost, and low energy requirements [26]. In addition, nanoadsorbents are characterized by their high specific surface area, chemical/thermal stability, and affinity for organic pollutants [27]. Although the efficiency of nanoadsorbents in the removal of organic compounds is remarkable, there are still limitations to conventional materials’ use, such as separation from the aqueous medium and the reuse of nanoadsorbents and nanocatalysts [28]. Recently, the development of nanocomposites as nanoadsorbents has been the subject of diverse research due to their increased surface area and physicochemical stability. Moreover, magnetic NCs have been used as a good alternative to improve the stability, textural properties, and reuse of nanoadsorbents [29]. The facilities separate material from the aqueous medium and considerably increase their reuse, resulting in high adsorptive capacity [30]. Additionally, the same behavior is observed for magnetic NCs as nanocatalysts. Using magnetic nanocatalysts allows the reuse of the material, increasing the cost-effectiveness and avoiding subsequent steps such as filtration and centrifugation [31].

Glyphosate {N-phosphomethyl[glycine] or (C₃H₈NO₅P)}, is an organophosphorus compound with herbicide properties discovered in 1970. It is a competitive inhibitor of the 5-enolpyruvylshikimate-3-phosphate synthase, an enzyme involved in aromatic amino acid biosynthesis in plants and microorganisms [32]. Glyphosate is now the most used herbicide globally, and its usage keeps increasing with the emergence of weed resistance, from 16 million kg spread in the world in 1994 to 79 million kg spread in 2014, including 15% in the United States alone [33]. Once in the environment, glyphosate is metabolized by microorganisms into aminomethylphosphonic acid (AMPA; known as its most active metabolite) and methylphosphonic acid (MPA) (Figure 1) [34]. Glyphosate and its metabolite AMPA can be found in soils, water, plants, food, and animals [35-37]. Glyphosate is detected in human urine, blood, and maternal milk, with urinary levels of 0.26-73.5 μg/l in exposed workers and 0.16-7.6 μg/l in the general population [38,39]. Glyphosate most likely enters the body via the dermal, oral and pulmonary routes [40]. Even if the dermal route allows a poor absorption (≈2%), it is the main reported route of entry in exposed farmers [41]. Glyphosate then seems to accumulate principally in the kidneys, liver, colon, and small intestine and is eliminated in the feces (90%) and urine within 48 h. Because of this omnipresence, its safety is of grave concern. Glyphosate has long been regarded as harmless allegedly because it targets an enzyme inexistent in animals, is supposedly degraded into CO₂, and its formulation contains misleadingly-called “inert” ingredients. Nevertheless, there is growing literature that describes the risks for glyphosate and glyphosate-based herbicides on human health [42]. After more than 40 years of global use, glyphosate has been classified as “probably carcinogenic” in humans by the International Agency for Research on Cancer (IARC). In March 2015, the World Health Organization’s IARC classified three organophosphates (glyphosate, malathion, and diazinon) as “probably carcinogenic for humans” (Category 2A) [43]. In contrast, in November 2015 the European Food Safety Agency determined glyphosate was “unlikely to pose a cancer risk for man” [44]. In 2018 the European Chemicals Agency, Risk Assessment Committee concluded that “the scientific evidence so far available does not satisfy the criteria for classifying glyphosate as carcinogenic, mutagenic or toxic for reproduction” [45]. In 2019, US federal health agency, the Agency for Toxic Substances and Disease Registry (ATSDR) [46], part of the Centers for Disease Control and Prevention [47], determined that both cancer and non-cancer hazards derive from exposure to glyphosate and glyphosate-based herbicides.

Figure 1: Glyphosate and main glyphosate by-products; aminomethylphosphonic acid (AMPA), methylphosphonic acid (MPA) and glyoxylate, respectively.

In modern agriculture, especially in most intensive and large-scale crops, herbicides are used to eliminate weeds. Glyphosate is a non-selective, highly effective, broad-spectrum, and low toxicity herbicide, whose usage increases exponentially for the effective in eliminating weeds indiscriminately [48]. In recent years, the long half-life of glyphosate and its main metabolite AMPA causes the existence in the environment. The potential impact of glyphosate in the environment is an increasing concern around the world. In the recent past, a significant increase in the use of the glyphosate herbicide has been noticed which further increased after the introduction of glyphosate-tolerant crops [49,50]. According to a report, The United States saw a 14 times increase in glyphosate use between 1992 and 2015, where the majority was applied to soybean and corn crops [51]. Being a nonselective, mutagenic, and carcinogenic herbicide, their presence in atmosphere can cause severe health and environmental issues [52]. As a result, it poses a high environmental risk and requires prompt studies towards its elimination.

In recent years, photocatalytic studies have explored the fabrication of ternary heterojunctions as a preferred scientific and practical method to improve the migration of photogenerated charge carriers [53]. Towards this end, ternary type II heterojunctions have shown significant success with accelerated charge carrier production [54]. However, the repulsion between the photogenerated electrons and the formation of weaker redox potentials limit its photocatalytic activity [55]. Therefore, another promising photosystem known as a Z-scheme heterojunction was developed to overcome the aforementioned issues [56]. In the case of Z-scheme photosystems, the conduction band electrons with a lower energy of one semiconductor migrate towards the valence band holes with a higher energy of other semiconductors. This combination leads to the formation of highly reductive electrons as well as highly oxidative holes [57]. Additionally, Z-scheme photo-systems not only enhance the charge separation efficiency of semiconductor photocatalysts but also possess electrons and holes with strong redox potential for superior photocatalytic applications. Moreover, ternary heterojunctions with a double electron transfer Z-scheme have photogenerated charge carriers with a prolonged lifetime compared to binary systems which improves the scope of light-harvesting [58]. Some recent ternary heterojunctions with double electron transfer Z-scheme channelization are g-C₃N₄/ZnO/ZnWO₄, polyaniline-BiOBr-GO, g-C₃N₄/Zn₂SnO₄N/ZnO [59-61].

The fabrication of NCs in ionic liquid (IL) media can provide a better scaling up approach for microscopic dispersion of particles and close interface contact between the individual components [62]. ILs as synthetic media provide unique advantages like negligible vapor pressure, thermal stability, and better conductivity than NCs. Additionally, the effect of “cation-π” and “π-π” interactions due to the presence of ionic liquids improve the nanoparticles (NPs) dispersion and stabilization which boosts the surface to volume ratio of NCs [63]. Recently, ionic liquids have been also employed for extensive polymerization and catalysis applications. Pahonik et al. [64] verified the oxidative polymerization of aniline with ammonium persulphate and the IL 1-butyl-3-methylimidazolium chloride (BMIMCl) under acidic conditions. More interestingly, IL-assisted NCs synthesis processes are relatively rapid, facile, greener, and more efficient without the requirement of any foreign stabilizer and surfactants [65]. The development of nanostructures and NCs with a simplified and greener IL-assisted in situ oxidative polymerization method is highly preferable method for photocatalytic degradation process of environmental pollutants.

Polyaniline (PANI) is a conducting polymer and organic semiconductor of the semi-flexible rod polymer family. PANI is one of the most studied conducting polymers [66,67]. To fabricate a ternary heterojunction, PANI can serve as the third active component of the photosystem. PANI has high demand as a low-cost and environment-friendly conjugated semiconductor for the fabrication of visible light harvesting photocatalysts [68]. It is a conducting polymer with an extensive conjugated π-system and high absorption coefficient towards visible light mediated charge carrier production. Furthermore, advantages like simple processing and high conductivity make it an emerging material for the synthesis of heterojunction materials. Recently, many efforts have been made to maximize the photo-harvesting efficiency of PANI-based composite materials. Researchers explored many positive hybrid effects arising from such systems due to the close contact of the interfaces of individual components leading to high separation efficiency of photogenerated electron-hole pairs [69].

The two-dimensional (2D) g-C₃N₄ semiconductor has a wide range of applications in the environmental and energy fields because of its visible-light activity, unique physicochemical properties, excellent chemical stability and low-cost [70,71]. Some important limitations of the photocatalytic activity of g-C₃N₄ are its low specific surface area, fast recombination of electrons and holes and poor visible light absorption [72-74]. To improve the above problems, the construction of a heterojunction with a suitable band gap semiconductor (co-catalyst) has been shown to be a good strategy to improve the photocatalytic performance of g-C₃N₄, such as g-C₃N₄-based conventional type II heterostructures, g-C₃N₄-based Z-scheme heterostructures, and g-C₃N₄-based p-n heterostructures, etc. The unique “Z” shape as the transport pathway of photogenerated charge carriers in Z-scheme photocatalytic systems is the most similar system to mimic natural photosynthesis in the many g-C₃N₄-based heterojunction photocatalysts. The construction of Z-scheme photocatalytic systems can promote visible light utilization and carrier separation, and maintain the strong reducibility and oxidizability of semiconductors [75-78]. There are many studies on g-C₃N₄-based Z-scheme heterojunction photocatalysts, such as ZnO/g-C₃N₄ [79-82], WO₃/g-C₃N₄ [83], g-C₃N₄/ZnS, g-C₃N₄/NiFe₂O₄ [84], g-C₃N₄/graphene/NiFe₂O₄ [85], NiCo/ZnO/g-C₃N₄ [86] and Bi₂Zr₂O₇/g-C₃N₄/Ag₃PO₄ [87], respectively. g-C₃N₄-based Z-scheme heterojunction photocatalysts have been made to improve the photocatalytic activity by combining with other semiconductor materials. Therefore, there are some problems with the single photocatalytic method, such as low adsorption ability, limited active sites and low removal efficiency. The integration of the adsorption and photocatalytic degradation of various organic pollutants is considered as a suitable and promising technology. On the other hand, it is still essential to fabricate photocatalysts with superior adsorption and degradation efficiencies.

g-C₃N₄ has been gaining great attention as a potential photocatalyst due to its stability and safety characteristics, as well as the fact that it can be facilely synthesized from low-cost raw materials. The low bandgap (~2.7 eV) can drive photo-oxidation reactions even under visible light [88-90]. However, the pure g-C₃N₄ has some drawbacks such as its low redox potential and high rate of recombination between photo-induced electrons and holes, which dramatically limits its photocatalytic efficiency. Several strategies have been investigated, including modification of the material’s size and structure [91], nonmetal and metal doping [92,93], and coupling with other photocatalysts [94-97]. For example, Liu et al. improved bulk g-C₃N₄’s performance in terms of Rhodamine B degradation from 30% to 100% by synthesizing mesoporous g-C₃N₄ nanorods through the nano-confined thermal condensation method. Dai et al. doped g-C₃N₄ with Cu through a thermal polymerization route and acquired a degradation rate of 90.5% with norfloxacin antibiotic. Nithya and Ayyappan, synthesized hybridized g-C₃N₄/ZnBi₂O₄ for reduction of 4-nitrophenol and reached an optimal removal efficiency of 79%. Among all, the construction of heterostructure photocatalysts by coupling g-C₃N₄ with other semiconductors seems to be an effective strategy to prevent electron and hole recombination, hence improving photocatalytic efficiency for contaminant treatment.

Zinc tungsten oxide or zinc tungstate (ZnWO₄) has received wide attention owing to its high ultraviolet (UV) light response, tunable band edges, optical transparency, easy availability, chemical stability, and adequate strength [98]. The band edge tunning of ZnWO₄-centered nanostructures can be organized through appropriate changes such as heterostructure construction, doped/combining with transition metal ions, and noble metals [99,100]. The alteration of electronic environment in ZnWO₄ nanomaterials through such engineered modifications can lead to interesting catalytic properties. The relationship between their structures and properties should therefore be considered to progress extremely proficient solar light conserving photocatalysts for the removal of toxic contaminants [101]. In the case of heterogeneous photocatalysis such as ZnWO₄, solid catalysts/semiconductors are utilized to remove organic pollutants under light irradiation due to redox reactions in photogenerated charge carriers. The mechanism is divided into three significant steps, generation of charge carrier pairs under irradiation, photogenerated charge carriers migrating on the surface of the catalyst, and initiation of the redox reaction by oxidative (OH^●) and superoxide (O₂^{– ●}) radicals [102]. For instance, Alshehri et al. [103] investigated that ZnWO₄ was used as a photocatalyst to degrade MB dye, and they reported the formation of OH^● and O₂^{– ●} free radicals oxidized the dye molecules to form inorganic minerals. The organic molecules by the photogenerated electron holes can also occur while hydroperoxyl radicals (OOH^●) and H₂O₂ are produced by the subsequent reactions, which occur between O₂^{– ●} and H⁺. This heterogeneous photocatalytic process induces the mineralization of organic pollutants (CO₂ and H₂O). Depending on the process’s efficiency, the pollutant’s composition, and its structure, additional products such as acids and salts can be formed. Exploration into photocatalysis has shown how UV-light, visible light, and solar irradiation can be utilized effectively to reduce environmental pollution [104,105]. Electron-hole pairs are produced when photon energy more prominent than the band gap of the semiconductor used to illuminate the semiconductor; this then leads to the formation of electron-hole pairs. OH^● when the generated electrons and holes react with H₂O and molecular oxygen on the surface of the crystal. With oxygen ions deposited around the tungsten, ZnWO₄ forms an insulated [WO₆] octahedron coordination with an asymmetric shape showing its local atomic structures with a monoclinic wolframite-type structure with the space group P2/c [106]. This is an essential inorganic ternary oxide material as it has been known to crystallize as a scheelite structure depending on the ionic radius [107]. However, W clusters form a network because they are more stable, which leads to forming the covalent nature of W-O bonds. In forming electron-hole pairs associated with a charge separation process and dipoles, the WO₆ clusters act as electron receptors. Thus, the oxygen vacancies in the Zn/W clusters can transfer electrons to the tungsten cluster and thus form permanent dipoles [108]. The Zn and W vacancies act as hole traps because they are negatively charged [109]. During the UV irradiation of ZnWO₄, the conduction band electrons generated are transferred to Ag nanocrystallite due to the Schottky barrier at Ag/ZnWO₄, which aid the charge carrier separation [110,111]. Different researchers have provided detailed and in-depth information, including improving ZnWO₄ as the next-generation catalysts for wastewater treatment. For instance, Gouveia et al. demonstrated that the overall performance of ZnWO₄ NPs was linked to the exposed surfaces of materials, their functional properties, and morphological structures; however, the authors failed to explain the concept of binary and multiple doping effects of ZnWO₄. According to the first-principle approach, the photocatalytic activity of ZnWO₄ depends on the intrinsic atomic properties and the electronic structure of the incomplete surface clusters of the exposed surfaces of the morphology. The authors found that the surface clusters in the morphology controlled the intrinsic atomic properties of the metal oxide in question. Furthermore, Geetha et al. [112] prepared ZnWO₄ nanoparticles via the co-precipitation method for the photocatalytic degradation of methylene blue (MB). The highest dye removal (81%) was observed for ZnWO₄ NPs prepared with 30 cm³ distilled water. Also, the performance of ZnWO₄ depended on the volume of the solvent (30-90 ml) and band gap energy (3.19 eV-3.16 eV), which was evidence of reduced interaction between metal and oxygen orbital. The members of the tungstate family have, over the years, been used for the mineralization of organic pollutants under UV [113] and sunlight [114] irradiation. However, the photocatalytic strength of ZnWO₄ stand-alone is not strong enough (Rahmani and Sedaghat, 2019). The enhancement of the photocatalytic activity of semiconductor ZnWO₄ for practical applications has deeply been considered for the degradation of contaminants. This has been the goal of many industries and scientists interested in environmental pollution control. However, a couple of approaches have been reported to further increase the properties of ZnWO₄ NPs for wastewater treatment.

In this study, a novel PANI/g-C₃N₄/ZnWO₄ ternary NCs as a heterostructure photocatalys was examined during photocatalytic degradation process in the efficient removal of glyphosate herbicide from a aqueous solution. Different pH values (3.0, 5.0, 7.0, 9.0 and 11.0), increasing glyphosate concentrations (5 mg/l, 10 mg/l, 15 mg/l and 20 mg/l), increasing PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations (5 mg/l, 15 mg/l, 30 mg/l and 45 mg/l) and increasing recycle times (1., 2., 3., 4., 5., 6. and 7.) was operated during photocatalytic degradation process in the efficient removal of glyphosate in a aqueous solution. The characteristics of the synthesized NPs were assessed using XRD, FESEM, EDX, FTIR, TEM, DRS and XPS analyses, respectively. The cyctotoxicity test was operated to the standard TBE (trypan blue dye exclusion) assay technique with Drosophila melanogaster (fruit fly). ANOVA statistical analysis was used for all experimental samples.

Materıals and Methods

Preparation of Graphitic Carbon Nitride (g-C₃N₄) Nanoparticles

g-C₃N₄ nanoparticles (NPs) was prepared by calcination of melamine (C₃H₆N₆) in a crucible with a lid at 550°C for 4 h. The obtained yellow powder was ground in an agate mortar after being cooled down to 25°C room temperature.

Preparation of Zinc Tungstate (ZnWO₄) Nanoparticles

ZnWO₄ NPs was prepared to sol-gel methods. Sol-gel method also called chemical solution deposition; it entails hydrolysis and polycondensation, gelation, aging, drying, densification, and crystallization. It is a highly effective method for synthesizing ZnWO₄ NPs with modified surfaces. Grossin [115] describe this method as involving the hydrolysis of the precursor in acidic or basic mediums and the polycondensation of the hydrolyzed. Rahmani and Sedaghat studied the nature of the ZnWO₄ NPs obtained from this study. The ZnWO₄ NPs were synthesized by adding 30 ml ethanol and 3 ml HCl into a mixture of zinc acetate dropwise, while sodium tungstate in deionized water was added to 20 ml of ethanol in a dropwise form. Both solutions were mixed vigorously, after which urea was added to the zinc acetate and sodium tungstate mixture. The ZnWO₄ NPs synthesized were characterized as well, where it was observed that the band gap energy was 3.20 eV. The ZnWO₄ NPs synthesized had an average diameter of between 26-78 nm.

Preparation of A Novel PANI/g-C₃N₄/ZnWO₄) (PGZ) Ternary Nanocomposites (NCs)

The novel PANI/g-C₃N₄/ZnWO₄) (PGZ) ternary NCs was synthesized by adopting an ionic liquid-assisted in situ oxidative polymerization process. The process includes the 1-butyl-3-methylimidazolium chloride-assisted polymerization of aniline using (NH₄)₂S₂O₈ as an oxidant. Firstly, 0.5 ml, 1.0 ml and 2.0 ml of aniline was added to an aqueous solution of 1-butyl-3-methylimidazolium chloride to make three different PANI solutions. Afterward, to each PANI mixture, an appropriate amount of (NH₄)₂S₂O₈ was added in a (NH₄)₂S₂O₈ /aniline=1/1 molar ratio. In two other round bottom flasks, 0.20 g g-C₃N₄ and 0.20 g ZnWO₄ were dispersed in 30 ml of 0.10 M HCl solution under ultra-sonication for 30 min. Subsequently, the particle mixtures were poured into the previously prepared PANI mixtures. The polymerization process was maintained for 12 h under mechanical stirring at 25°C room temperature. The products were separated by centrifugation and washed multiple times with ethanol to remove the residual IL media. The three composite mixtures obtained were dried in a vacuum oven at 70°C. The prepared samples are marked as xPGZ (0.5-PGZ, 1-PGZ, and 2-PGZ), where x denotes the amount of aniline added. For simplicity, sample 1-PGZ is referred to as PGZ throughout this study.

Photocatalytic Degradation Reactor

A 2 liter cylinder quartz glass reactor was used for the photodegradation experiments in the glyphosate aqueous solution at different operational conditions. 1000 ml glyphosate aqueous solution was filled for experimental studies and the photocatalyst were added to the cylinder quartz glass reactors. The UV-A lamps were placed to the outside of the photo-reactor with a distance of 3 mm. The photocatalytic reactor was operated with constant stirring (1.5 rpm) during the photocatalytic degradation process. 10 ml of the reacting solution were sampled and centrifugated (at 10000 rpm) at different time intervals. The UV irradiation treatments were created using one or three UV-A lamp emitting in the 350-400 nm range (λ_max=368 nm; FWHM=17 nm; Actinic BL TL-D 18W, Philips). Three 50 W UV-A lamps (Total: 150 W UV-A lamps) were used during experimental conditions for this study.

Glyphosate Photocatalytic Degradation Experiments

The photocatalytic degradation efficiencies of PANI, g-C₃N₄ NCs, ZnWO₄ NCs and PANI/g-C₃N₄/ZnWO₄ ternary NCs were investigated with a cylinder quartz glass photocatalytic reactor under UV-vis light irradiation. The series of glyphosate degradation studies were performed in an aqueous solution. The temperature of the photocatalytic system was maintained using continuously circulating aqueous solution. Typically, 25 mg/l catalysts were used for the batch degradation study with 100 ml of 10 mg/l glyphosate under continuous magnetic stirring. The pH=7.0 ± 0.1 of the pollutant solutions was maintained throughout the degradation process by adding H₂SO₄ and NaOH solutions as necessary. Initially, the reaction mixtures were kept in dark to check the adsorption properties of glyphosate and to attain adsorption-desorption equilibrium. Next, the whole setup was exposed to UV-vis light for the photocatalytic degradation study. In 20 min time gap, a 4 ml aliquot of the pollutant solution was withdrawn and centrifuged to separate the NCs. The initial and final concentration supernatants were analyzed using a UV-vis spectrometer for detection of the intermediates and degradation products formed during the photocatalytic degradation process. The percentage degradation was calculated by the following Equation (1):

Determination of Glyphosate and Photodegradation by-Products

The quantification of glyphosate and glyphosate major photodegradation products was determined to a Gas Chromatography-Mass Spectrometry (GC-MS). These samples were performed with a gas chromatographya gas chromatographically (Agilent 6890N GC) equipped with a mass selective detector (Agilent 5973 inert MSD) (GC-MS) (Hewlett-Packard 6980/HP5973MSD). A capillary column (HP5-MS, 30 m, 0.25 mm, 0.25 m) was used. The initial oven temperature was kept at 50°C for 1 min, then raised to 200°C at 25°C/min and from 200°C to 300°C at 8°C/min, and then maintained for 5.5 min. High purity He(g) was used as the carrier gas at constant flow mode (1.5 ml/min, 45 cm/s linear velocity). The method involves the addition of 5% borate buffer to the aqueous sample to adjust the pH=9.0 and then mixing with 9-fluorenylmethyl chloroformate (FMOC) in acetonitrile prior to analysis. The derivatization process was continued for 16 h at 25°C and the process was stopped by drop-wise addition of 6 M HCl solution where the resulting pH was measured to be pH=1.5. Chromatographic separation was performed with a C18 column where the mobile phase was 5 mM HAc/NH4Ac (pH=4.8) acetonitrile. The acetonitrile percentage was changed from 75% (0-42 min) to 100% (42.1-45 min) to 5% (45.1-50 min). For each sample separation process was completed in 50 min. The degradation products were detected at 210 nm with a PDA detector. The same method was also applied for derivatization and analysis of glyphosate and glyphosate by-products; acetate, aminomethylphosphonic acid (AMPA), phosphate, sarcosine and glycine as standards.

Quantification of Major Oxygen Species

To quantify the reactive oxygen species (OH^● and O₂^{– ●}) production under light illumination, 1.2 g/l benzoic acid and 5×10⁻⁵ mol/l nitro blue tetrazolium dichloride (NBT) solutions were considered as molecular probes, respectively. 100 mg of PANI/g-C₃N₄/ZnWO₄ ternary NCs was dispersed in 100 ml of the molecular probe solutions to evaluate the radical production efficiency. For every 10 min, 3 ml of sample was pipetted out for further analysis. The NBT sample was analyzed with a UV spectrometer (Shimadzu 2450) at 258 nm. The quantification of O₂^{– ●} was done by the NBT degradation method. The quantity of OH^● was measured by analyzing the amount of p-hydroxybenzoic acid in the sample with the same GC-MS method mentioned above. In this case, the mobile phase was acetonitrile/water (30/70) with a 1 ml/min flow rate.

Characterization

X-Ray Diffraction Analysis

Powder XRD patterns were recorded on a Shimadzu XRD-7000, Japan diffractometer using Cu Kα radiation (λ=1.5418 Å, 40 kV, 40 mA) at a scanning speed of 1°/min in the 10-80° 2θ range. Raman spectrum was collected with a Horiba Jobin Yvon-Labram HR UV-Visible NIR (200-1600 nm) Raman microscope spectrometer, using a laser with the wavelength of 512 nm. The spectrum was collected from 10 scans at a resolution of 2 /cm. The zeta potential was measured with a SurPASS Electrokinetic Analyzer (Austria) with a clamping cell at 300 mbar.

Field Emission Scanning Electron Microscopy (FESEM) and Energy Dispersive X-Ray (EDX) Spectroscopy Analysis

The morphological features and structure of the synthesized catalyst were investigated by FESEM (FESEM, Hitachi S-4700), equipped with an EDX spectrometry device (TESCAN Co., Model III MIRA) to investigate the composition of the elements present in the synthesized catalyst.

Fourier Transform Infrared Spectroscopy (FTIR) Analysis

The FTIR spectra of samples was recorded using the FT-NIR spectroscope (RAYLEIGH, WQF-510).

Transmission Electron Microscopy (TEM) Analysis

The structure of the samples were analysed TEM analysis. TEM analysis was recorded in a JEOL JEM 2100F, Japan under 200 kV accelerating voltage. Samples were prepared by applying one drop of the suspended material in ethanol onto a carbon-coated copper TEM grid, and allowing them to dry at 25°C room temperature.

Diffuse Reflectance UV-Vis Spectra (DRS) Analysis

DRS Analysis in the range of 200-800 nm were recorded on a Cary 5000 UV-Vis Spectrophotometer from Varian. DRS was used to monitor the glyphosate concentration in experimental samples.

X-Ray Photoelectron Spectroscopy (XPS) Analysis

The valence state of the biogenic palladium nanoparticles was investigated and was analyzed using XPS (ESCALAB 250Xi, England). XPS used an Al Ka source and surface chemical composition and reduction state analyses was done, with the core levels recorded using a pass energy of 30 eV (resolution ≈0.10 eV). The peak fitting of the individual core-levels was done using XPS-peak 41 software, achieving better fitting and component identification. All binding energies were calibrated to the C 1s peak originating from C-H or C-C groups at 284.6 eV.

Cytotoxicity Test

The standard TBE (trypan blue dye exclusion) assay technique was followed for to check the cytotoxicity of the photo-treated glyphosate solution and photocatalyst. Drosophila melanogaster (fruit fly) was considered as a model organism since 75% of its disease genome sequence is functionally homologous to that of humans [116]. Similar studies were also reported where Drosophila melanogaster was employed as a model research organism to study the toxic effects of various chemicals, drugs, medicines and NPs. The TBE assay has been studied for differentiating live and dead cells in the Drosophila melanogaster larval gut. Before the cytotoxicity study, glyphosate samples (untreated, 5 mg/l, 10 mg/l, 15 mg/l and 20 mg/l) were prepared. And, to analyze the cytotoxic effects of both the initial glyphosate solution and phototreated products, the TBE assay was implemented. Firstly, third instar larvae were taken and washed with 1× PBS to remove food particles that remained in the larval body. Then, the larvae were transferred to a Petri plate with 2% solidified agar to keep them hungry. After that, 10 3rd instar larvae were transferred to 1.5 ml eppendorf tubes each containing 500 μl of the glyphosate solutions, respectively. These larvae were kept for 30 min to feed on the chemical orally. After the incubation, the larvae were washed once with 1× PBS. Then, the larvae were transferred into a container with 0.5% TBE solution and kept for 45 min in a dark atmosphere at 25°C room temperature. After incubation, again the excess strain was washed twice with 1× PBS for 10 min each [117]. Then, further analysis was done with a USB stereomicroscope and digital images were taken to check any abnormality in the gut. A similar procedure was followed for different amounts of the PANI/g-C₃N₄/ZnWO₄ ternary NCs samples (5 mg, 15 mg, 30 mg and 45 mg). Finally, the percentage of the defective Drosophila melanogaster larva is calculated as following Equation (2):

Statistical Analysis

ANOVA analysis of variance between experimental data was performed to detect F and P values. The ANOVA test was used to test the differences between dependent and independent groups [118]. Comparison between the actual variation of the experimental data averages and standard deviation is expressed in terms of F ratio. F is equal (found variation of the date averages/expected variation of the date averages). P reports the significance level, and d.f indicates the number of degrees of freedom. Regression analysis was applied to the experimental data in order to determine the regression coefficient R² [119]. The aforementioned test was performed using Microsoft Excel Program.

All experiments were carried out three times and the results are given as the means of triplicate samplings. The data relevant to the individual pollutant parameters are given as the mean with standard deviation (SD) values.

Results and Discussions

A Novel PANI/g-C₃N₄/ZnWO₄ Ternary NCs Characteristics

The Results of X-Ray Diffraction (XRD) Analysis

The results of XRD analysis was observed to pure g-C₃N₄ NPs, pure ZnWO₄ NPs, pure PANI and PANI/g-C₃N₄/ZnWO₄ ternary NCs, respectively, in aqueos solution with photocatalytic degradation process for glyphosate removal (Figure 2). The characterization peaks were observed at 2θ values of 12.71° and 28.84°, respectively, corresponding to the (100) and (002) planes of implying pure g-C₃N₄ NPs in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 2a). The characterization peaks were obtained at 2θ values of 17.10°, 19.52°, 14.70°, 15.01°, 30.17°, 37.28°, 39.11°, 42.34°, 44.41°, 46.53°, 49.20°, 50.65°, 53.42°, 54.41°, 61.36°, 65.22°, and 68.74°, respectively, corresponding to the (010), (100), (011), (110), (111), (021), (200), (121), (112), (211), (002), (220), (130), (202), (032), (311) and (041), respectively, implying pure ZnWO₄ NPs in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 2b). The characterization peaks were found at 2θ values of 18.27°, 24.32°, 26.11°, 28.44°, 30.10°, 37.22°, 41.34°, 53.28°, 61.34° and 64.42°, respectively, corresponding to (100), (011), (110), (002), (111), (021), (200), (121), (202), (032) and (311), respectively, implying PANI in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 2c). The characterization peaks were observed at 2θ values of 28.39°, 30.17°, 37.63°, 41.20°, 54.33°, 61.20° and 64.60°, respectively, corresponding to (002), (111), (021), (121), (202), (033) and (312), respectively, implying PANI in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 2d).

Figure 2: The XRD patterns of (a) pure g-C₃N₄ NPs, (b) pure ZnWO₄ NPs, (c) PANI and (d) PANI/g-C₃N₄/ZnWO₄ ternary NCs, respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Results of Field Emission Scanning Electron Microscopy (FESEM) Analysis

The morphological features of pure g-C₃N₄ NPs, pure ZnWO₄ NPs, PANI and PANI/g-C₃N₄/ZnWO₄ ternary NCs were characterized through FE-SEM images (Figure 3). The FESEM images of pure g-C₃N₄ NPs were obtained in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 3a). The FESEM images of pure ZnWO₄ NPs were observed in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 3b). The FESEM images of PANI were viewed in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 3c). The FESEM images of PANI/g-C₃N₄/ZnWO₄ ternary NCs were characterized in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 3d).

Figure 3: FESEM images of (a) pure g-C₃N₄ NPs, (b) pure ZnWO₄ NPs, (c) PANI and (d) PANI/g-C₃N₄/ZnWO₄ NCs, respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Results of Energy Dispersive X-Ray (EDX) Spectroscopy Analysis

The EDX analysis was also performed to investigate the composition of pure g-C₃N₄ NPs (Figure 4a), pure ZnWO₄ NPs (Figure 4b), PANI (Figure 4c) and PANI/g-C₃N₄/ZnWO₄ NCs (Figure 4d), respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

Figure 4: EDX images of (a) pure g-C₃N₄ NPs, (b) pure ZnWO₄ NPs, (c) PANI and (d) PANI/g-C₃N₄/ZnWO₄ ternary NCs, respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Results of Fourier Transform Infrared Spectroscopy (FTIR) Analysis

The FTIR spectrum of pure g-C₃N₄ NPs, pure ZnWO₄ NPs, PANI and PANI/g-C₃N₄/ZnWO₄ ternary NCs, respectively, were determined in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 5). The main peaks of FTIR spectrum for pure ZnWO₄ NPs (black spectrum) was observed at 3421 1/cm, 1326 1/cm, 1015 1/cm and 678 1/cm wavenumber, respectively (Figure 5a). The main peaks of FTIR spectrum for pure g-C₃N₄ NPs (green spectrum) was obtained at 3348 1/cm, 1645 1/cm, 1410 1/cm, 1234 1/cm and 807 1/cm wavenumber, respectively (Figure 5b). The main peaks of FTIR spectrum for PANI (blue spectrum) was determined at 3151 1/cm, 1544 1/cm, 1408 1/cm, 1239 1/cm, 900 1/cm and 815 1/cm wavenumber, respectively (Figure 5c). The main peaks of FTIR spectrum for PANI/g-C₃N₄/ZnWO₄ ternary NCs (red spectrum) was obtained at 3416 1/cm, 1638 1/cm, 1074 1/cm and 549 1/cm wavenumber, respectively (Figure 5d).

Figure 5: FTIR spectrum of (a) pure ZnWO₄ (black spectrum), (b) g-C₃N₄ NPs (green spectrum), (c) PANI (blue spectrum) and (d) PANI/g-C₃N₄/ZnWO₄ ternary NCs (red spectrum), respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Results of Transmission Electron Microscopy (TEM) Analysis

The TEM images of PANI/g-C₃N₄/ZnWO₄ ternary NCs was observed in micromorphological structure level in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 6).

Figure 6: TEM images of PANI/g-C₃N₄/ZnWO₄ ternary NCs in micromorphological structure level in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Results of Diffuse reflectance UV-Vis Spectra (DRS) Analysis

The absorption spectra of glyphosate was observed in DRS Analysis (Figure 7). First, the absorption spectra of glyphosate were obtained at a maximum concentration of 15 mg/l in the wavelength range from 300 nm to 800 nm using diffuse reflectance UV-Vis spectra (Figure 7). Absorption peaks were observed at wavelengths of 375 nm for pure g-C₃N₄ NPs (red pattern) (Figure 7a), 390 nm for pure ZnWO₄ NPs (blue pattern) (Figure 7b), 370 nm for PANI (green patern) (Figure 7c) and 430 nm for PANI/g-C₃N₄/ZnWO₄ NCs (black pattern) (Figure 7d), respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

Figure 7: The DRS patterns of (a) pure g-C₃N₄ NPs (red pattern) (b) pure ZnWO₄ NPs (blue pattern), (c) PANI (green pattern) and (d) PANI/g-C₃N₄/ZnWO₄ NCs (black pattern), respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Results of X-Ray Photoelectron Spectroscopy (XPS) Analysis

The XPS analysis of pure g-C₃N₄ NPs, pure ZnWO₄ NPs, PANI and PANI/g-C₃N₄/ZnWO₄ ternary NCs, respectively, were perforned to investigate in aqueous solution with photocatalytic degradation process for glyphosate removal (Figure 8). Absorption peaks were observed at binding energy of 401.51 eV for pure g-C₃N₄ NPs (blue pattern) (Figure 8a), 399.63 eV for pure ZnWO₄ NPs (green pattern) (Figure 8b), 398.12 eV for PANI (red patern) (Figure 8c) and 398.36 eV for PANI/g-C₃N₄/ZnWO₄ NCs (black pattern) (Figure 8d), respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

Figure 8: The XPS spectra of (a) pure g-C₃N₄ NPs (blue pattern) (b) pure ZnWO₄ NPs (green pattern), (c) PANI (red pattern) and (d) PANI/g-C₃N₄/ZnWO₄ NCs (black pattern), respectively, in aqueous solution with photocatalytic degradation process for glyphosate removal.

The Reaction Kinetics of Glyphosate Herbicide

The reaction kinetics glyphosate were investigated using the Langmuir-Hinshelwood first-order kinetic model, expressed by Eddy et al. [119], as following Equation (3):

where; r_o: denotes the initial photocatalytic degradation reaction rate (mg/l.min), and k: denotes the rate constant of a first-order reaction. At the beginning of the reaction, t=0, C_t=C₀, the equation can be obtained after integration as following Equation (4):

where; C₀ and C: are the initial and final concentration (mg/l) of glyphosate; the solution at t (min) and k (1/min) are the rate constant.

The pollutants photocatalytic degradation rate was found using a pseudo first-order reaction kinetic equation (Equation 5):

where; K_app: is the apparent rate constant, C₀: is the pollutant concentration before illumination and C_t: is the final concentration of the pollutant at time t.

The correlation coefficients had R² values greater than 0.9, as a result, the first-order kinetic model fit the experimental data well. The first-order rate constants (k) were determined from the slope of the linear plots.

Photocatalytic Degradation Mechanisms

The possible photocatalytic reactions for glyphosate degradation over the PANI/g-C₃N₄/ZnWO₄ ternary heterojunction (PGZ) can be expressed as following Equation (6), Equation (7), Equation (8), Equation (9), Equation (10), Equation (11), Equation (12) and Equation (13):

The photocatalytic degradation mechanism can be better understood when it is correlated to the kinetics of the degradation reaction. The rate constants were determined from the equation ln(C_t/C₀)=K_appt, where K_app is the apparent rate constant for the reaction, and C₀ and C_t represent the initial and final (after time t) concentrations of glyphosate. The apparent rate constants were calculated from the experimental data. Linear fitting between the experimental data and pseudo-first order kinetic model suggested that the degradation process of glyphosate follows the pseudo-first-order kinetic model. The optimized results indicate the highest photo-degradation ability and kinetics for the PANI/g-C₃N₄/ZnWO₄ ternary NCs, which may be due to its suitable composition and enhanced surface active sites as suggested by BET (Brunner-Emmett-Teller) analysis.

Effect of Increasing pH values for Glyphosate Removal in Aqueous Solution during Photocatalytic Degradation Process

Increasing pH values (pH=3.0, pH=5.0, pH=7.0, pH=9.0 and pH=11.0, respectively) was examined during photocatalytic degradation process in aqueous solution for glyphosate removal (Figure 9). 42%, 58%, 71% and 89% glyphosate removal efficiencies was measured at pH=3.0, pH=5.0, pH=7.0 and pH=9.0, respectively, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at 25°C (Figure 9). The maximum 99% glyphosate removal efficiency was obtained during photocatalytic degradation process in aqueous solution, at pH=11.0, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time and at 25°C, respectively (Figure 9).

Figure 9: Effect of increasing pH values for glyphosate removal in aqueous solution during photocatalytic degradation process, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time and at 25°C, respectively.

Effect of Increasing Glyphosate Concentrations for Glyphosate Removal in Aqueous Solution during Photocatalytic Degradation Process

Increasing glyphosate concentrations (5 mg/l, 10 mg/l, 15 mg/l and 20 mg/l) were operated at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0, at 25°C, respectively (Figure 10). 60%, 85% and 73% glyphosate removal efficiencies were obtained to 5 mg/l, 10 mg/l and 20 mg/l glyphosate concentrations, respectively, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C (Figure 10). The maximum 99% glyphosate removal efficieny was found with photocatalytic degradation process in aqueous solution, at 15 mg/l glyphosate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively (Figure 10).

Figure 10: Effect of increasing glyphosate concentrations for glyphosate removal in aqueous solution during photocatalytic degradation process, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively.

Effect of Increasing PANI/g-C₃N₄/ZnWO₄ Ternary NCs Concentrations for Glyphosate Removals in Aqueous Solution during Photocatalytic Degradation Process

Increasing PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations (5 mg/l, 15 mg/l, 30 mg/l and 45 mg/l) were operated at 15 mg/l glyphosate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0, at 25°C, respectively (Figure 11). 51%, 75% and 82% glyphosate removal efficiencies were obtained to 5 mg/l, 15 mg/l and 45 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations, respectively, at 15 mg/l glyphossate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0, at 25°C, respectively (Figure 11). The maximum 99% glyphosate removal efficieny was measured to 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs with photocatalytic degradation process in aqueous solution, at 15 mg/l glyphosate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively (Figure 11).

Figure 11: Effect of increasing PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations for glyphosate removal in aqueous solution during photocatalytic degradation process, at 15 mg/l glyphosate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively.

The Results of Cytotoxicity Test

The cytotoxicity of PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst and the glyphosate solutions were tested with the TBE assay analytical protocol and by considering with Drosophila melanogaster larvae before and after photocatalytic degradation process. Cytotoxicity test was performed with untreated glyphosate solution and after photodegradation process sample of different glyphosate concentrations (5 mg/l, 10 mg/l, 15 mg/l and 20 mg/l) and different PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations (5 mg/l, 15 mg/l, 30 mg/l and 45 mg/l), at 25°C, at pH=7.0, respectively (Table 1).

98%, 95%, 90% and 80% cyctotoxicity removal efficiencies were obtained to 5 mg/l, 10 mg/l, 15 mg/l and 20 mg/l glyphosate concentrations, respectively, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively (Table 1). The maximum 99% cyctotoxicity removal was observed at untreated glyphosate samples, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively (Table 1).

96%, 82% and 74% cyctotoxicity removal efficiencies were measured to 15 mg/l, 30 mg/l and 45 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst concentrations, respectively, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively (Table 1). The maximum 99% cyctotoxicity removal was observed at 5 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst concentrations, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively (Table 1). The study revealed the excellent minimization of cytotoxicity of glyphosate after photocatalytic degradation process with the PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst. Also, the PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst is found to be non-cytotoxic irrespective of its quantity used.

Table 1: Effect of increasing glyphosate and PANI/g-C₃N₄/ZnWO₄ ternary NCs concentrations on cyctotoxicity test in aqueous solution after photocatalytic degradation process, at 25°C, at pH=7.0, respectively.

Effect of Different Recycle Times for Glyphosate Removals in Aqueous Solution during Photocatalytic Degradation Process

Different recycle times (1., 2., 3., 4., 5., 6. and 7.) were operated for glyphosate removals in aqueous solution during photocatalytic degradation process, at 15 mg/l glyphosate, 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively (Figure 12). 92%, 87%, 84%, 80%, 76% and 73% glyphosate removal efficiencies were measured after 2. recycle time, 3. recycle time, 4. recycle time, 5. recycle time, 6. recycle time and 7. recycle time, respectively, at 15 mg/l glyphosate, 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively (Figure 12). The maximum 99% glyphosate removal efficiency was measured in aqueous solution during photocatalytic degradation process, after 1. recycle time, at 15 mg/l glyphosate, 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively (Figure 12).

Figure 12: Effect of recycle times for glyphosate removal in aqueous solution during photocatalytic degradation process, at 15 mg/l glyphosate, 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively.

Conclusıons

The maximum 99% glyphosate removal efficiency was obtained during photocatalytic degradation process in aqueous solution, at pH=11.0, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time and at 25°C, respectively.The maximum 99% glyphosate removal efficieny was found with photocatalytic degradation process in aqueous solution, at 15 mg/l glyphosate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively.The maximum 99% glyphosate removal efficieny was measured to 30 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs with photocatalytic degradation process in aqueous solution, at 15 mg/l glyphosate, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=11.0 and at 25°C, respectively.The maximum 99% cyctotoxicity removal was observed at untreated glyphosate samples, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively. The maximum 99% cyctotoxicity removal was observed at 5 mg/l PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst concentrations, after 180 min photocatalytic degradation time, at 150 W UV-vis light irradiation power, at pH=7.0 and at 25°C, respectively. The study revealed the excellent minimization of cytotoxicity of glyphosate after photocatalytic degradation process with the PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst. Also, the PANI/g-C₃N₄/ZnWO₄ ternary NCs photocatalyst is found to be non-cytotoxic irrespective of its quantity used.

As a result, the a novel PANI/g-C₃N₄/CoMoO₄ ternary NCs photocatalyst during photocatalytic degradation process in aqueous solution for glyphosate removal was stable in harsh environments such as acidic, alkaline, saline, and then was still effective process. When the amount of contaminant was increased, the a novel PANI/g-C₃N₄/CoMoO₄ ternary NCs photocatalyst during photocatalytic degradation process performance was still considerable. The synthesis and optimization of a novel PANI/g-C₃N₄/CoMoO₄ ternary NCs heterostructure photocatalyst provides insights into the effects of preparation conditions on the material’s characteristics and performance, as well as the application of the effectively designed photocatalyst in the removal of gylphosate herbicites, which can potentially be deployed for purifying wastewater, especially agricultural industry wastewater treatment. Finally, the combination of a simple, easy operation preparation process, excellent performance and cost effective, makes this a novel PANI/g-C₃N₄/CoMoO₄ ternary NCs heterostructure photocatalyst a promising option during photocatalytic degradation process in agricultural industry wastewater treatment.

Acknowledgement

This research study was undertaken in the Environmental Microbiology Laboratories at Dokuz Eylül University Engineering Faculty Environmental Engineering Department, Izmir, Turkey. The authors would like to thank this body for providing financial support.

References

1. Rout PR, Zhang TC, Bhunia P, Surampalli RY (2021) Treatment technologies for emerging contaminants in wastewater treatment plants: A review. Sci Total Environ 753: 141990. [crossref]
2. Song B, Xu P, Chen M, Tang W, Zeng G, et al. (2019) Using nanomaterials to facilitate the phytoremediation of contaminated soil Crit Rev Environ Sci Tecnol 49: 791-824.
3. Xiang H, Zhang Y, Atkinson D, Sekar R (2020) Effects of anthropogenic subsidy and glyphosate on macroinvertebrates in streams. Environ Sci Pollut Res 27: 21939-21952.[crossref]
4. García-García CR, Parrón T, Requena M, Alarcón R, Tsatsakis AM, et al. (2016) Occupational pesticide exposure and adverse health effects at the clinical, hematological and biochemical level. Life Sci145: 274-283. [crossref]
5. Saleh IA, Zouari N, Al-Ghouti MA (2020) Removal of pesticides from water and wastewater: Chemical, physical and biological treatment approaches. Environ Technol Innov 19: 101026.
6. Sharma A, Kumar V, Shahzad B, Tanveer M, Sidhu GPS, et al. (2019) Worldwide pesticide usage and its impacts on ecosystem. Sn Appl Sci 11: 1446-1462.
7. Mostafalou S, Abdollahi M (2013) Pesticides and human chronic diseases: Evidences, mechanisms, and perspectives. Toxicol Appl Pharmacol 268: 157-177.[crossref]
8. Casida JE, Durkin KA (2017) Pesticide chemical research in toxicology: Lessons from nature. Chem Res Toxicol 30: 94-104. [crossref]
9. Abdennouri M, Baâlala M, Galadi A, El Makhfouk M, Bensitel M, et al. (2016). Photocatalytic degradation of pesticides by titanium dioxide and titanium pillared purified clays. Arab J Chem 9: S313-S318.
10. Liu G, Li L, Huang X, Zheng S, Xu X, et al. (2018) Adsorption and removal of organophosphorus pesticides from environmental water and soil samples by using magnetic multi-walled carbon nanotubes@ organic framework ZIF-8. J Mater Sci 53: 10772-10783.
11. Pandiselvam R, Kaavya R, Jayanath Y, Veenuttranon K, Lueprasitsakul P, et al. (2020) Ozone as a novel emerging technology for the dissipation of pesticide residues in foods—A review. Trends Food Sci Technol 97: 38-54.
12. Barka N, Qourzal S, Assabbane A, Nounah A, Ait-Ichou Y (2010) Photocatalytic degradation of an azo reactive dye, Reactive Yellow 84, in water using an industrial titanium dioxide coated media. Arab J Chem 3: 279-283.
13. Dehghani MH, Karamitabar Y, Changani F, Heidarinejad Z (2019) High performance degradation of phenol from aqueous media using ozonation process and zinc oxide nanoparticles as a semiconductor photo catalyst in the presence of ultraviolet radiation. Desalin Water Treat 166: 105-114.
14. Ibhadon AO, Fitzpatrick P (2013) Heterogeneous photocatalysis: Recent advances and applications. Catalysts 3: 189-218.
15. Silva WL, Lansarin MA, Livotto PR, Santos JHZ (2015) Photocatalytic degradation of drugs by supported titania-based catalysts produced from petrochemical plant residue. Powder Technol 279: 166-172.
16. Dariani RS, Esmaeili A, Mortezaali A, Dehghanpour S (2016) Photocatalytic reaction and degradation of methylene blue on TiO₂ nano-sized particles. Optik 127: 7143-7154.
17. Kamat PV (2012) TiO₂ Nanostructures: Recent Physical Chemistry Advances. J Phys Chem C 116: 11849-11851.
18. Hernández S, Hidalgo D, Sacco A, Chiodoni A, Lamberti A, et al. (2015). Comparison of photocatalytic and transport properties of TiO₂ and ZnO nanostructures for solar-driven water splitting. Phys Chem Chem Phys 17: 7775-7786.[crossref]
19. Das S, Srivastava VC (2017) Synthesis and characterization of ZnO/CuO nanocomposite by electrochemical method. Mater Sci Semicond Process 57: 173-177.
20. Saravanan R, Karthikeyan S, Gupta VK, Sekaran G, Narayanan V, et al. (2013) Enhanced photocatalytic activity of ZnO/CuO nanocomposite for the degradation of textile dye on visible light illumination. Mater Sci Eng C 33: 91-98. [crossref]
21. Da Silva Bruckmann F, Ledur CM, da Silva IZ, Dotto GL, Rhoden CRB (2022) A DFT theoretical and experimental study about tetracycline adsorption onto magnetic graphene oxide. J Mol Liq 353: 118837.
22. Rane AV, Kanny K, Abitha VK, Thomas S (2018) Methods for synthesis of nanoparticles and fabrication of nanocomposites. In Synthesis of Inorganic Nanomaterials; Woodhead Publishing: Amsterdam 121-139.
23. Omanovi´c-Mikliˇcanin E, Badnjevi´c A, Kazlagi´c A, Hajlovac M (2020) Nanocomposites: A brief review. Health Technol 10: 51-59.
24. Nunes FB, Da Silva Bruckmann F, Da Rosa Salles T, Rhoden CBR (2022) Study of phenobarbital removal from the aqueous solutions employing magnetite-functionalized chitosan. Environ Sci Pollut Res 12: 908-931. [crossref]
25. Khan I; Saeed K; Khan I (2019) Nanoparticles: Properties, applications and toxicities. Arab J Chem 12: 908-931.
26. Yoldi M, Fuentes-Ordoñez E, Korili S, Gil A (2019) Zeolite synthesis from industrial wastes. Microporous Mesoporous Materials 287: 183-191.
27. Mintova S, Jaber M, Valtchev V (2015) Nanosized microporous crystals: Emerging applications. Chem Soc Rev 44: 7207-7233.[crossref]
28. Nguyen CH, Tran HN, Fu CC, Lu YT, Juang RS (2020) Roles of adsorption and photocatalysis in removing organic pollutants from water by activated carbon-supported titania composites: Kinetic aspects. J Taiwan Inst Chem Eng 109: 51-61.
29. Bruckmann FS, Rossato Viana A, Tonel MZ, Fagan SB, Garcia WJS, et al. (2022) Influence of magnetite incorporation into chitosan on the adsorption of the methotrexate and in vitro cytotoxicity. Environ Sci Pollut Res 29: 70413-70434.[crossref]
30. Rahimi B, Jafari N, Abdolahnejad A, Farrokhzadeh H, Ebrahimi A (2019) Application of efficient photocatalytic process using a novel BiVO/TiO₂-NaY zeolite composite for removal of acid orange 10 dye in aqueous solutions: Modeling by response surface methodology (RSM) J Environ Chem Eng 7: 103253.
31. Falyouna O, Eljamal O, Maamoun I, Tahara A, Sugihara Y (2020) Magnetic zeolite synthesis for efficient removal of cesium in a lab-scale continuous treatment system. J Colloid Interface Sci 571: 66-79.[crossref]
32. Steinrucken H, Amrhein N (1984) 5-Enolpyruvylshikimate-3-phosphate syn- € thase of Klebsiella pneumoniae 2. Inhibition by glyphosate [N-(phosphonomethyl)glycine]. Eur J Biochem 143: 351-357.[crossref]
33. Benbrook C (2016) Trends in glyphosate herbicide use in the United States and globally. Environ Sci Eur 28: Article number: 3.[crossref]
34. Williams G, Kroes R, Munro I (2000) Safety evaluation and risk assessment of the herbicide Roundup and its active ingredient, glyphosate, for humans. Regul Toxicol Pharmacol 31: 117-165.[crossref]
35. Alferness P, Iwata Y (2002) Determination of glyphosate and (aminomethyl)- phosphonic acid in soil, plant and animal matrixes, and water by capillary gas chromatography with mass-selective detection. Am Chem Soc 42: 2751-2759.
36. Caloni F, Cortinovis C, Rivolta M, Davanzo F (2016) Suspected poisoning of domestic animals by pesticides. Sci Total Environ 539: 331-336.[crossref]
37. El-Gendy K, Mosallam E, Ahmed N, Aly N (2018) Determination of glyphosate residues in Egyptian soil samples. Anal Biochem 557: 1-6.[crossref]
38. Acquavella J, Alexander B, Mandel J, Gustin C, Baker B, et al. (2004) Glyphosate biomonitoring for farmers and their families: Results from the Farm Family Exposure Study. Environ Health Perspect 112: 321-326. [crossref]
39. Gillezeau C, van Gerwen M, Shaffer R, Rana I, Zhang L, et al. (2019) The evidence of human exposure to glyphosate: A review. Environ Health 18: 2.
40. Martinez T, Long W, Hiller, R (1990) Comparison of the toxicology of the herbicide Roundup by oral and pulmonary routes of exposure. Proc West Pharmacol Soc 33: 193-197.[crossref]
41. Connolly A, Coggins M, Galea K, Jones K, Kenny L, et al. (2019) Evaluating glyphosate exposure routes and their contribution to total body burden: A study among amenity horticulturalists. Ann Work Expo Health 63: 133-147.[crossref]
42. Vandenberg LN, Blumberg B, Antoniou MN, Benbrook CM, Carroll L, et al. (2017) Is it time to reassess current safety standards for glyphosate-based herbicides? J Epidemiol Community Health 71: 613-618.
43. IARC (2015) Outdoor Air Pollution. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans 109.
44. EFSA (2016) The 2014 European Union Report on pesticide residues in food. EFSA J 14: e04611.
45. ATSDR (2019) Toxicological Profile for Silica. Atlanta, GA: U.S. Department of Health and Human Services, Public Health Service.
46. CDC (2009) Fourth National Report on Human Exposure to Environmental Chemicals. Atlanta (GA)
47. Huang Y, Li Z, Yao K, Chen C, Deng C, et al. (2021) Suppressing toxic intermediates during photocatalytic degradation of glyphosate by controlling adsorption modes. Appli Catal B: Environ 299: 120671.
48. Duke SO, Powles SB (2009) Glyphosate-resistant crops and weeds: now and in the future. AgBioForum 12: 346-357.
49. Annett R, Habibi HR, Hontela A (2014) Impact of glyphosate and glyphosate-based herbicides on the freshwater environment. J Appl Toxicol 34: 458-479. [crossref]
50. United States Geological Survey, Pesticide National Synthesis Project, 2014.
51. Torretta V, Katsoyiannis IA, Viotti P, Rada EC (2018) Critical review of the effects of glyphosate exposure to the environment and humans through the food supply chain. Sustainability 10: 950.
52. Wang H, Zhang L, Chen Z, Hu J, Li S, et al. (2014) Semiconductor heterojunction photocatalysts: design, construction, and photocatalytic performances. Chem Soc Rev 43: 5234-5244.
53. Cai H, Wang B, Xiong L, Bi J, Yuan L, et al. (2019) Orienting the charge transfer path of type-II heterojunction for photocatalytic hydrogen evolution. Appl Catal B 256: 117853.
54. Di T, Xu Q, Ho W, Tang H, Xiang Q, et al. (2019) Review on metal sulphide-based Z-scheme photocatalysts. ChemCatChem 11: 1394-1411.
55. Low J, Jiang C, Chen B, Wageh S, Al-Ghamdi AA, et al. (2017) A review of direct Z-scheme photocatalysts. Small Methods 1: 1700080.
56. Yang C, Xue Z, Qin J, Sawangphruk M, Rajendran S, et al. (2019) Visible-light driven photocatalytic H₂ generation and mechanism insights on Bi₂O₂CO₃/G-C₃N₄ Z-scheme photocatalyst. J Phys Chem C 123: 4795-4804.
57. Liang C, Niu CG, Guo H, Huang DW, Wen XJ, et al. (2018) Combination of efficient charge separation with the assistance of novel dual Z-scheme system: self-assembly photocatalyst Ag@AgI/BiOI modified oxygen-doped carbon nitride nanosheet with enhanced photocatalytic performance. Catal Sci Technol 8: 1161-1175.
58. Zhu L, Li H, Xu Q, Xiong D, Xia P (2020) High-efficient separation of photoinduced carriers on double Z-scheme heterojunction for superior photocatalytic CO₂ J Colloid Interface Sci 564: 303-312.[crossref]
59. Liu X, Cai L (2019) A novel double Z-scheme BiOBr-GO-polyaniline photocatalyst: Study on the excellent photocatalytic performance and photocatalytic mechanism. Appl Surf Sci 483: 875-887.
60. Wang M, Tan G, Ren H, Xia A, Liu Y (2019) Direct double Z-scheme O-g-C₃N₄/Zn₂SnO₄N/ZnO ternary heterojunction photocatalyst with enhanced visible photocatalytic activity. Appl Surf Sci 492: 690-702.
61. Ravula S, Zhang C, Essner JB, Robertson JD, Lin J, et al. (2017) Ionic Liquid-Assisted Synthesis of Nanoscale (MoS₂)_x(SnO₂)_1-x on Reduced Graphene Oxide for the Electrocatalytic Hydrogen Evolution Reaction. ACS Appl Mater Interfaces, 9: 8065-8074.[crossref]
62. Barik B, Kumar A, Nayak PS, Achary LSK, Rout L, et al. (2020) Ionic liquid assisted mesoporous silica-graphene oxide nanocomposite synthesis and its application for removal of heavy metal ions from water. Materials Chemistry and Physics, 239: 122028.
63. Pahovnik D, Žagar E, Vohlídal J, Žigon M (2010) Ionic liquid-induced formation of polyaniline nanostructures during the chemical polymerization of aniline in an acidic aqueous medium. Synthetic Metals 160: 1761-1766.
64. Zhao S, Fang J, Wang Y, Zhang Y, Zhou Y (2019) Poly(ionic liquid)-assisted synthesis of open-ended carbon nitride tube for efficient photocatalytic hydrogen evolution under visible-light irradiation. ACS Sustainable Chem Eng 7: 10095-10104.
65. Okamoto Y, Brenner W (1964) Organic Semiconductors. New York: Reinhold Publishing Corporation.
66. Heeger, A (201) Nobel Lecture: Semiconducting and metallic polymers: The fourth generation of polymeric materials. Reviews of Modern Physics 73: 681-700.
67. Chen S, Huang D, Zeng G, Xue W, Lei L, et al. (2020) In-situ synthesis of facet-dependent BiVO₄/Ag₃PO₄/PANI photocatalyst with enhanced visible-light-induced photocatalytic degradation performance: Synergism of interfacial coupling and hole-transfer. Chem Eng Sci 382: 122840.
68. Jiang J, Li H, Zhang L (2012) New insight into daylight photocatalysis of AgBr@Ag: synergistic effect between semiconductor photocatalysis and plasmonic photocatalysis. Chemistry A European Journal 18: 6360-6369.
69. Fu J, Yu J, Jiang C, Cheng B (2018) g-C₃N₄-based heterostructured photocatalysts. Adv Energy Mater 8: 1701503.
70. Ren Y, Zeng D, Ong WJ (2019) Interfacial engineering of graphitic carbon nitride (g-C₃N₄)-based metal sulfide heterojunction photocatalysts for energy conversion: A review, Chin J Catal 40: 289-319.
71. Hao Q, Chen T, Wang R, Feng J, Chen D (2018a) A separation-free polyacrylamide/bentonite/graphitic carbon nitride hydrogel with excellent performance in water treatment. J Clean Prod 197: 1222-1230.
72. Hao X, Zhou J, Cui Z, Wang Y, Wang Y, et al. (2018b) Zn-vacancy mediated electron-hole separation in ZnS/g-C₃N₄ heterojunction for efficient visible-light photocatalytic hydrogen production. Appl Catal B 229: 41-51.
73. Mestre AS, Carvalho AP (2019) Photocatalytic degradation of pharmaceuticals carbamazepine, diclofenac, and sulfamethoxazole by semiconductor and carbon materials: a review, Molecules 24: 3702. [crossref]
74. Chen Y, Fan ZX, Zhang ZC, Niu WX, Li CL, et al. (2018) Two-dimensional metal nanomaterials: synthesis, properties, and applications. Chem Rev 118: 6409-6455.
75. Xu Q, Zhang L, Yu J, Wageh S, Al-Ghamdi AA, et al. (2018) Direct Z-scheme photocatalysts: Principles, synthesis, and applications. Mater Today 21: 1042-1063.
76. Huang D, Chen S, Zeng G, Gong X, Zhou CY, et al. (2019) Artificial Z-scheme photocatalytic system: What have been done and where to go?, Coord Chem Rev, 385: 44-80.
77. Zhang X, Lou S, Zeng Y (2020) Facile fabrication of a novel visible light active g-C₃N₄-CoMoO₄ heterojunction with largely improved photocatalytic performance. Mater Lett 281: 128661.
78. Yu W, Xu D, Peng T (2015) Enhanced photocatalytic activity of g-C₃N₄ for selective CO₂ reduction to CH₃OH via facile coupling of ZnO: a direct Z-scheme mechanism. J Mater Chem A 3: 19936-19947.
79. Liu X, Pang F, He M, Ge J (2017a) Confined reaction inside nanotubes: New approach to mesoporous g-C₃N₄ Nano Res 10: 3638-3647.
80. Liu D, Zhang M, Xie W, Sun L, Chen Y, et al. (2017b) Porous BN/TiO2 hybrid nanosheets as highly efficient visible-light-driven photocatalysts, Appl Catal B-Environ 207: 72-78.
81. Liu S, Chen F, Li S, Peng X, Xiong Y (2017) Enhanced photocatalytic conversion of greenhouse gas CO₂ into solar fuels over g-C₃N₄ nanotubes with decorated transparent ZIF-8 nanoclusters. Appl Catal B 211: 1-10.
82. Cui L, Ding X, Wang Y, Shi H, Huang L, et al. (2017). Facile preparation of Z-scheme WO₃/g-C₃N₄ composite photocatalyst with enhanced photocatalytic performance under visible light. Appl Surf Sci B 391: 202-210.
83. Gebreslassie G, Bharali P, Chandra U, Sergawie A, Boruah PK, et al. (2019a) Novel g-C₃N₄/graphene/NiFe₂O₄ nanocomposites as magnetically separable visible light driven photocatalysts. J Photochem Photobiol A 382: 111960.
84. Gebreslassie G, Bharali P, Chandra U, Sergawie A, Boruah PK, et al. (2019b) Hydrothermal synthesis of g-C₃N₄/NiFe₂O₄ nanocomposite and its enhanced photocatalytic activity. Appl Organomet Chem 33: e5002.
85. Wu J, Hu J, Qian H, Li J, Yang R, et al. (2022) NiCo/ZnO/g-C₃N₄ Z-scheme heterojunction nanoparticles with enhanced photocatalytic degradation oxytetracycline. Diamond Relat Mater 121: 108738.
86. Qu Z, Jing Z, Chen X, Wang Z, Ren H, et al. (2023) Preparation and photocatalytic performance study of dual Z-scheme Bi₂Zr₂O₇/g-C₃N₄/Ag₃PO₄ for removal of antibiotics by visible-light. J Environ Sci 125: 349-361. [crossref]
87. Zhao H, Yu H, Quan X, Chen S, Zhang Y., et al. (2014) Fabrication of atomic single layer graphitic C₃N₄ and its high performance of photocatalytic disinfection under visible light irradiation. Appl Catal B 152-153: 46-50.
88. Mamba G. Mishra AK (2016) Graphitic carbon nitride (g-C₃N₄) nanocomposites: A new and exciting generation of visible light driven photocatalysts for environmental pollution remediation. Appl Catal B 198: 347-377.
89. Zhang C, Li Y, Shuai D, Shen Y, Xiong W, et al. (2019) Graphitic carbon nitride (g-C₃N₄)-based photocatalysts for water disinfection and microbial control: A review, Chemosphere 214: 462-479.[crossref]
90. Darkwah WK, Ao Y (2018) Mini review on the structure and properties (photocatalysis), and preparation techniques of graphitic carbon nitride nano-based particle, and its applications. Nanoscale Res Lett 13: 388. [crossref]
91. Oh WD, Chang VWC, Hu ZT, Goei R, Lim TT (2017) Enhancing the catalytic activity of g-C₃N₄ through Me doping (Me = Cu, Co and Fe) for selective sulfathiazole degradation via redox-based advanced oxidation process Chem Eng J 323: 260-269.
92. Dai Y, Gu Y, Bu Y (2020) Modulation of the photocatalytic performance of g-C₃N₄ by two-sites co-doping using variable valence metal. Appl Surf Sci 500: 144036.
93. Ye L, Liu J, Jiang Z, Peng T, Zan L (2013) Facets coupling of BiOBr-g-C₃N₄ composite photocatalyst for enhanced visible-light-driven photocatalytic activity. Appl Catal B 142-143: 1-7.
94. Tian N, Huang H, He Y, Guo Y, Zhang T, et al. (2015) Mediator-free direct Z-scheme photocatalytic system: BiVO₄/g-C₃N₄ organic-inorganic hybrid photocatalyst with highly efficient visible-light-induced photocatalytic activity, Dalton Trans., 44: 4297-4307.[crossref]
95. Nithya R, Ayyappan S (2020) Novel exfoliated graphitic C₃N₄ hybridised ZnBi₂O₄ (g-C₃N₄/ZnBi₂O₄) nanorods for catalytic reduction of 4-Nitrophenol and its antibacterial activity. J Photochem Photobiol Chem 398: 112591.
96. Tian N, Huang H, Wang S, Zhang T, Du X, et al. (2020) Facet-charge-induced coupling dependent interfacial photocharge separation: A case of BiOI/g-C₃N₄ p-n junction. Appl Catal B 267: 118697.
97. Reddy CV, Koutavarapu R, Reddy IN, Shim J (2021) Effect of a novel onedimensional zinc tungsten oxide nanorods anchored two-dimensional graphitic carbon nitride nanosheets for improved solar-light-driven photocatalytic removal of toxic pollutants and photoelectrochemical water splitting. J Mater Sci Mater Electron 32: 33-46.
98. Ojha DP, Kim HJ (2020) Investigation of photocatalytic activity of ZnO promoted hydrothermally synthesized ZnWO4 nanorods in UV-visible light irradiation. Chem Eng Sci 212: 115338.
99. Ke J, Younis MA, Kong Y, Zhou H, Liu J, et al. (2018) Nanostructured ternary metal tungstate-based photocatalysts for environmental purification and solar water splitting: a review. Nano-Micro Lett 10: 69.[crossref]
100. Liu H, Huang J, Chen J, Zhong J, Li J, et al. (2019) Preparation and characterization of novel Ag/Ag₂WO₄/ZnWO₄ heterojunctions with significantly enhanced sunlight-driven photocatalytic performance. Solid State Sci 95: 105923.
101. Enesca A (2021) The influence of photocatalytic reactors design and operating parameters on the wastewater organic pollutants removal – a mini review. Catalysts 11: 556: 1-22.
102. Alshehri SM, Ahmed J, Ahamad T, Alhokbany N, Arunachalam P., et al. (2018) Synthesis, characterization, multifunctional electrochemical (OGR/ ORR/SCs) and photodegradable activities of ZnWO₄ J Sol Gel Sci Technol 87: 137-146.
103. Zhang L, Wang Z, Wang L, Xing Y, Li X, et al. (2014a) Electrochemical performance of ZnWO₄/CNTs composite as anode materials for lithium-ion battery. Appl Surf Sci 305: 179-185.
104. Zhang C, Zhang H, Zhang K, Li X, Leng Q, et al. (2014b) Photocatalytic activity of ZnWO₄: band structure, morphology and surface modification. ACS Appl Mater Interfaces 6: 14423-14432.
105. He G, Fan H, Ma L, Wang K, Ding D, et al. (2016) Synthesis, characterization and optical properties of nanostructured ZnWO₄. Mater Sci Semicond Process 41: 404-410.
106. Chena G, Wanga F, Yu J, Zhang H, Zhang X (2017) Improved red emission by codoping Li⁺³ in ZnWO₄: Eu⁺³ J Mol Struct 1128: 1-4.
107. Gouveia AF, Assis M, Cavalcante LS, Gracia L, Longo E, et al. (2018) Reading at exposed surfaces: theoretical insights into photocatalytic activity of ZnWO₄. Front Res Today 1: 1005.
108. Zhai B, Yang L, Huang Y (2019) Intrinsic defect engineering in Eu⁺³ doped ZnWO₄ for annealing temperature tunable photoluminescence. Nanomaterials 9: 99.
109. Li M, Zhu Q, Li JG, Kim BN (2020a) Elongation of ZnWO₄ nanocrystals for enhanced photocatalysis and the effects of Ag decoration. Appl Surf Sci 515: 146011.
110. Li S, Chang L, Peng J, Gao J, Lu J, et al. (2020b) Bi⁰ nanoparticle loaded on Bi₃þ-doped ZnWO₄ nanorods with oxygen vacancies for enhanced photocatalytic NO removal. J Alloys Compd 818: 52837.
111. Geetha GV, Keerthana SP, Madhuri K, Sivakumar R (2021) Effect of solvent volume on the properties of ZnWO₄ nanoparticles and their photocatalytic activity for the degradation of cationic dye. Inorg Chem Commun 132: 108810.
112. Rahmani M, Sedaghat T (2019) Nitrogen-doped ZnWO₄ nanophotocatalyst: synthesis, characterization and photodegradation of methylene blue under visible light. Res Chem Intermed 45: 5111-5124.
113. Paliki AK, Suresh P, Sailja VB (2016) Rapid visible light photocatalytic degradation of organic pollutants using ZnWO₄ Int J Eng Appl Sci Technol 1: 183.
114. Grossin D (2021) Encyclopedia of materials: technical ceramics and glasses. Phosphates 3: 567-574.
115. Bergman P, Esfahani SS, Engström Y (2017) Drosophila as a model for human diseases-focus on innate immunity in barrier epithelia. In: Current Topics in Developmental Biology 121: 29-81.[crossref]
116. Fernández Bedmar Z, Anter J, de La Cruz Ares S, Muñoz Serrano A, Alonso Moraga A, et al. (2011) Role of citrus juices and distinctive components in the modulation of degenerative processes: genotoxicity, antigenotoxicity, cytotoxicity, and longevity in Drosophila. J Toxicol Environ Health A 74: 1052-1066. [crossref]
117. Zar JH (1984) Biostatistical analysis, Prentice-Hall, Englewood Cliffs.
118. Statgraphics Centurion XV, sofware, StatPoint Inc, Statgraphics Centurion XV, Herndon, VA, USA, 2005.
119. Eddy DR, Puri FN, Noviyanti AR (2015) Synthesis and photocatalytic activity of silica-based sand quartz as the supporting TiO₂ Proced Chem 17: 55-58.

DOI: 10.31038/NAMS.2023621

Abstract

In this study, a novel graphitic carbon nitride/cobalt molybdate (g-C₃N₄/CeO₂) nanocomposites (NCs) as a photocatalys was examined during photocatalytic degradation process in the efficient removal of Ofloxacin (OFX) from pharmaceutical industry wastewater plant, İzmir, Turkey. Different pH values (3.0, 4.0, 6.0, 7.0, 9.0 and 11.0), increasing OFX concentrations (5 mg/l, 10 mg/l, 20 mg/l and 40 mg/l), increasing g-C₃N₄/CeO₂ NCs concentrations (1 mg/l, 2 mg/l, 4 mg/l, 6 mg/l, 8 mg/l and 10 mg/l), different g-C₃N₄/CeO₂ NCs mass ratios (5/5, 6/4, 7/3, 8/2, 9/1, 1/9, 2/8, 3/7 and 4/6), increasing recycle times (1., 2., 3., 4., 5., 6. and 7.) was operated during photocatalytic degradation process in the efficient removal of OFX in pharmaceutical industry wastewater. The characteristics of the synthesized nanoparticles (NPs) were assessed using X-Ray Difraction (XRD), Field Emission Scanning Electron Microscopy (FESEM), Energy-Dispersive X-Ray (EDX), Fourier Transform Infrared Spectroscopy (FTIR), Transmission Electron Microscopy (TEM), and Diffuse reflectance UV-Vis spectra (DRS) analyses, respectively. The acute toxicity assays were operated with Microtox (Aliivibrio fischeri also called Vibrio fischeri) and Daphnia magna acute toxicity tests. The photocatalytic degradation mechanisms of g-C₃N₄/CeO₂ NCs and the reaction kinetics of OFX were evaluated in pharmaceutical industry wastewater during photocatalytic degradation process. ANOVA statistical analysis was used for all experimental samples. The maximum 99% OFX removal efficiency was obtained during photocatalytic degradation process in pharmaceutical industry wastewater, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively. The maximum 99% OFX removal efficieny was found with photocatalytic degradation process in pharmaceutical industry wastewater, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min, at pH=6.0 and at 25°C, respectively. The maximum 99% OFX removal efficieny was measured to 8 mg/l g-C₃N₄/CeO₂ NCs with photocatalytic degradation process in pharmaceutical industry wastewater, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min, at pH=6.0 and at 25°C, respectively. The maximum 99% OFX removal efficiency was measured at 2/8wt g-C₃N₄/CeO₂ NCs mass ratios at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min, at pH=6.0 and at 25°C, respectively. The maximum 99% OFX removal efficiency was measured in pharmaceutical industry wastewater during photocatalytic degradation process, after 1. recycle time, at 20 mg/l OFX, 8 mg/l g-C₃N₄/CeO₂ NCs, at 2/8wt g-C₃N₄/CeO₂ NCs mass ratio, after 180 min, at pH=6.0 and at 25°C, respectively. 96.41% maximum Microtox (Aliivibrio fischeri) acute toxicity removal yield was found in OFX=20 mg/l after 180 min photocatalytic degradation time and at 60°C. It was observed an inhibition effect of OFX=40 mg/l to Microtox with Vibrio fischeri after 180 min and at 60°C. 92.38% maximum Daphnia magna acute toxicity removal was obtained in OFX=20 mg/l after 180 min photocatalytic degradation time and at 60°C, respectively. It was observed an inhibition effect of OFX=40 mg/l to Daphnia magna after 180 min and at 60°C. OFX concentrations > 20 mg/l decreased the acute toxicity removals by hindering the photocatalytic degradation process. Similarly, a significant contribution of increasing OFX concentrations to acute toxicity removal at 60°C after 180 min, was not observed. It can be concluded that the toxicity originating from the OFX is not significant and the real acute toxicity throughout photocatalytic degradation process was attributed to the pharmaceutical industry wastewater, to their metabolites and to the photocatalytic degradation process by-products. As a result, the a novel g-C₃N₄/CeO₂ NCs photocatalyst during photocatalytic degradation process in pharmaceutical industry wastewater was stable in harsh environments such as acidic, alkaline, saline, and then was still effective process. When the amount of contaminant was increased, the a novel g-C₃N₄/CeO₂ NCs photocatalys during photocatalytic degradation process performance was still considerable. The synthesis and optimization of g-C₃N₄/CeO₂ heterostructure photocatalyst provides insights into the effects of preparation conditions on the material’s characteristics and performance, as well as the application of the effectively designed photocatalyst in the removal of antibiotics, which can potentially be deployed for purifying wastewater, especially pharmaceutical wastewater. Finally, the combination of a simple, easy operation preparation process, excellent performance and cost effective, makes this a novel g-C₃N₄/CeO₂ NCs a promising option during photocatalytic degradation process in pharmaceutical industry wastewater treatment.

Keywords

ANOVA statistical analysis, Antibiotics, Coronavirus Disease-2019 (COVID-19), Cost analysis, Diffuse reflectance UV-Vis spectra (DRS), Electrochemical filtration process, Energy-dispersive X-ray (EDX), Field emission scanning electron microscopy (FESEM), Fourier transform infrared spectroscopy (FTIR), Hydrothermal-calcination method, Hydroxly (OH●) radicals, Microtox (Aliivibrio fischeri or Vibrio fischeri) and Daphnia magna acute toxicity tests, Nanoparticles (NPs), Novel graphitic carbon nitride/cerium dioxide nanocomposites (g-C3N4/CeO2 NCs), Ofloxacin (OFX), Pharmaceutical industry wastewater, Photocatalytic degradation mechanisms, Reaction kinetics, Sol–gel method, Transmission Electron Microscopy (TEM), Ultraviolet (UV), X-ray difraction (XRD)

Introductıon

Emerging contaminants (ECs), sometimes known as contaminants of emerging concern (CECs) can refer to a wide variety of artificial or naturally occurring chemicals or materials that are harmful to human health after long-term disclosure. ECs can be classified into several classes, including agricultural contaminants (pesticides and fertilizers), medicines and antidote drugs, industrial and consumer waste products, and personal care and household cleaning products [1,2]. Antibiotics are one of the ECs that have raised concerns in the previous two decades because they have been routinely and widely used in human and animal health care, resulting in widespread antibiotic residues discharged in surface, groundwater, and wastewater.

Antibiotics, which are widely utilized in medicine, poultry farming and food processing, have attracted considerable attention due to their abuse and their harmful effects on human health and the ecological environment. The misuse of antibiotics induces Deoxyribonucleic Acid (DNA) contamination and accelerates the generation of drug-resistant bacteria and super-bacteria thus, some diseases are more difficult to cure . A number of studies have revealed that the level of antibiotics in the soil, air and surface water, and even in potable water, is excessive in many areas, which will ultimately accumulate in the human body via drinking water and then damage the body’s nervous system, kidneys and blood system. Therefore, it is necessary to develop an efficient method to remove antibiotics present in pharmaceutical industry wastewater [3-13].

The uncontrolled, ever-growing accumulation of antibiotics and their residues in the environment is an acute modern problem. Their presence in water and soil is a potential hazard to the environment, humans, and other living beings. Many therapeutic agents are not completely metabolized, which leads to the penetration of active drug molecules into the biological environment, the emergence of new contamination sources, the wide spread of bacteria and microorganisms with multidrug resistance. Modern pharmaceutical wastewater facilities do not allow efficient removal of antibiotic residues from the environment, which leads to their accumulation in ecological systems . Global studies of river pollution with antibiotics have shown that 65% of surveyed rivers in 72 countries on 6 continents are contaminated with antibiotics. According to the World Health Organization (WHO), surface and groundwater, as well as partially treated water, containing antibiotics residue and other pharmaceuticals, typically at < 100 ng/l concentrations, whereas treated water has < 50 ng/l concentrations, respectively . However, the discovery of ECs in numerous natural freshwater sources worldwide is growing yearly. Several antibiotic residues have been reported to have been traced at concentrations greater than their ecotoxicity endpoints in the marine environment, specifically in Europe and Africa . Thus, the European Union’s Water Framework Directive enumerated certain antibiotics as priority contaminants. In some rivers, the concentrations were so high that they posed a real danger to both the ecosystem and human health. This matter, the development of effective approaches to the removal of antibiotics from the aquatic environment is of great importance [14-26].

The removal of antibiotics and their residues from water and wastewater prior to their final release into the environment is of particular concern. Modern purification methods can be roughly divided into the following three categories depending on the purification mechanism: biological treatment, chemical degradation, and physical removal. Each of these methods has its own advantages and disadvantages. For example, biological purification can remove most antibiotic residues, but the introduction of active organisms into the aquatic environment can upset the ecological balance. Various chemical approaches (ozonation, chlorination, and Fenton oxidation) cannot provide complete purification and, in some cases, lead to the death of beneficial microorganisms due to low selectivity. Photocatalysis is widely used in new environmental control strategies. However, this method has a number of key disadvantages, such as insufficient use of visible light, rapid annihilation of photogenerated carriers, and incomplete mineralization, which greatly limits its application [27-33].

Ofloxacin (OFX) is a quinolone antibiotic useful for the treatment of a number of bacterial infections . A quinolone antibiotic is a member of a large group of broad-spectrum bacteriocidals that share a bicyclic core structure related to the substance 4-quinolone . They are used in human and veterinary medicine to treat bacterial infections, as well as in animal husbandry, specifically poultry production . OFX is well-known for their antimicrobial and anti-inflammatory capabilities . OFX is used to treat pneumonia, skin and urinary tract infections . Severe acute respiratory syndrome (SARS)-CoV-2 (COVID-19) pandemic, which has killed and infected people in 216 countries/territories, has become the most significant pandemic of the century . OFX combined with other drugs, has been widely used to minimise COVID-19-induced inflammation in 2020.OFX is a typical fluoroquinolone antibiotic administered to both humans and animals, and after administration, approximately 78% of OFX is excreted. OFX pharmaceutical compounds enter water resources in various ways, such as human and animal excretions and inefficient industrial wastewater treatment. In the class of antibiotics, OFX is also recognised as highly refractory and persistent in aquatic water systems. As the biodegradation of OFX is difficult, sewage treatment plants (STPs) have a low removal rate, and the OFX concentrations in the STP effluents of Beijing, Hangzhou, and Vancouver have been determined to be between 6×10^-7 and 1.405×10^-3 mg/l [34-41].

Generally, the advanced oxidation processes (AOPs), such as the Fenton or Fenton-like reaction, ozonation or catalytic ozonation, photocatalytic oxidation, electrochemical oxidation, and ionizing radiation, have been widely used for antibiotics degradation in recent years . One of the most promising techniques applied for efficient degradation of antibiotics are Advanced Oxidation Processes (AOPs). Nowadays, particular attention is paid to photocatalytic reactions, in which highly oxidizing species responsible for mineralization of organic pollutants are formed in-situ in the reaction media by means of light and a photocatalyst . The photocatalytic activity is closely related to the physicochemical properties but also to the morphology and texture of the materials studied, for this reason the synthesis techniques are often of great importance. Photocatalysis, which occurs under exposure to UV light, is also a common method for the environmental pollutant elimination . The conventional photocatalysis utilizes mostly UV from sunlight, which accounts for only 4% of the solar energy. Therefore, through the introduction of catalysts, the utilization rate of sunlight can be effectively improved. To overcome the low-efficiency problem of the photocatalysis, the development of a more efficient catalyst system that would effectively improve the catalytic oxidation efficiency and overcome the existing limitations is important. The catalytic activity of the catalyst can be effectively improved by modulating its surface area, preparation method, and changing its properties and structures [42-57].

Numerous materials have been reported to have the potential and capacity to treat water or wastewater polluted with these antibiotics residue by applying the processes of adsorption and catalytic oxidation during the last few decades. The reported materials include mesoporous carbon beads, clay minerals, activated carbon, cellulose, and chitosan. As a result of engineering and science evolution, and in complement to the urgent need to increase the adsorption capability of antibiotic contaminants, more advanced materials such as carbon nanotube (CnT), nano-zero valent iron (nZVI), nanoporous carbons, porous graphene and graphene oxide (GO), to date have been analyzed and improved in their ability to remove these ECs from water [58-85].

Nanomaterials with a high specific surface area are a promising platform for the development and production of low-cost and highly efficient sorbents for various pollution molecules. For example, graphene-based nanomaterials were utilized to remove antibiotics, which are adsorbed on the material surfaces due to π-π-, electrostatic or hydrophobic interactions, as well as the formation of hydrogen bonds. Highly efficient antibiotic sorption was also observed when using highly porous, surface-active, and structurally stable silica-based materials, metal oxide NPs, and metal-organic frameworks. The photocatalysts, which mainly rely on the production of highly oxidizing species such as hydroxyl radical (OH^●) and superoxide anion radical (O₂^{− ●}), have been considered an effective approach for the degradation of antibiotics in water [86-103].

The two-dimensional (2D) g-C₃N₄ semiconductor has a wide range of applications in the environmental and energy fields because of its visible-light activity, unique physicochemical properties, excellent chemical stability and low-cost. Some important limitations of the photocatalytic activity of g-C₃N₄ are its low specific surface area, fast recombination of electrons and holes and poor visible light absorption. To improve the above problems, the construction of a heterojunction with a suitable band gap semiconductor (co-catalyst) has been shown to be a good strategy to improve the photocatalytic performance of g-C₃N₄, such as g-C₃N₄-based conventional type II heterostructures, g-C₃N₄-based Z-scheme heterostructures, and g-C₃N₄-based p–n heterostructures, etc. The unique “Z” shape as the transport pathway of photogenerated charge carriers in Z-scheme photocatalytic systems is the most similar system to mimic natural photosynthesis in the many g-C₃N₄-based heterojunction photocatalysts. The construction of Z-scheme photocatalytic systems can promote visible light utilization and carrier separation, and maintain the strong reducibility and oxidizability of semiconductors. There are many studies on g-C₃N₄-based Z-scheme heterojunction photocatalysts, such as ZnO/g-C₃N₄, WO₃/g-C₃N₄, g-C₃N₄/ZnS,, g-C₃N₄/NiFe₂O₄, g-C₃N₄/graphene/NiFe₂O₄, NiCo/ZnO/g-C₃N₄ and Bi₂Zr₂O₇/g-C₃N₄/Ag₃PO₄, respectively. g-C₃N₄-based Z-scheme heterojunction photocatalysts have been made to improve the photocatalytic activity by combining with other semiconductor materials. Therefore, there are some problems with the single photocatalytic method, such as low adsorption ability, limited active sites and low removal efficiency. The integration of the adsorption and photocatalytic degradation of various organic pollutants is considered as a suitable and promising technology. On the other hand, it is still essential to fabricate photocatalysts with superior adsorption and degradation efficiencies [104-121].

g-C₃N₄ has been gaining great attention as a potential photocatalyst due to its stability and safety characteristics, as well as the fact that it can be facilely synthesized from low-cost raw materials. The low bandgap (~2.7 eV) can drive photo-oxidation reactions even under visible light. However, the pure g-C₃N₄ has some drawbacks such as its low redox potential and high rate of recombination between photo-induced electrons and holes, which dramatically limits its photocatalytic efficiency. Several strategies have been investigated, including modification of the material’s size and structure, nonmetal and metal doping, and coupling with other photocatalysts. For example, Liu et al. improved bulk g-C₃N₄’s performance in terms of Rhodamine B degradation from 30% to 100% by synthesizing mesoporous g-C₃N₄ nanorods through the nano-confined thermal condensation method. Dai et al. doped g-C₃N₄ with Cu through a thermal polymerization route and acquired a degradation rate of 90.5% with norfloxacin antibiotic. Nithya and Ayyappan, synthesized hybridized g-C₃N₄/ZnBi₂O₄ for reduction of 4-nitrophenol and reached an optimal removal efficiency of 79%. Among all, the construction of heterostructure photocatalysts by coupling g-C₃N₄ with other semiconductors seems to be an effective strategy to prevent electron and hole recombination, hence improving photocatalytic efficiency for contaminant treatment [122-131].

CeO₂ (Ceria or Cerium(IV) oxide) is a versatile, inert, and physically and chemically stable material with multiple and diverse applications. Due to its hardness (Mohs scale 7), it was initially used as an abrasive material, but today it is used (alone or in binary or complex mixtures) in the field of heterogeneous catalysis (oxidation of hydrocarbons) or in the field of sensors, energy, and fuels such as solid oxide fuel cells, but also in water-splitting processes or photocatalysis. CeO₂ applications in the dermato-cosmetics industry and in the biomedical field (antibacterial effect) should also be mentioned here. CeO₂ is also possible to combine two or more properties, for example, the infrared filtering properties with the photocatalytic ones, to optimize practical applications. CeO₂ is semiconductor photocatalyst with various applications and similar properties to TiO₂. However, its band gap is in the wide range of 2.6 to 3.4 eV, depending on the preparation method. Furthermore, CeO₂ exhibits promising photocatalytic activity. Nonetheless, the position of CB and VB limits its application as an efficient photocatalyst utilizing solar energy, even though CeO₂ can absorb a larger fraction of the solar spectrum than TiO₂. The photocatalytic and photoelectrocatalytic activity of CeO₂ in wastewater treatment can be improved by various modification techniques, including changes in morphology, doping with metal cation dopants and non-metal dopants, coupling with other semiconductors, combining it with carbon supporting materials, etc.. The main properties that make CeO₂ significant as a photocatalyst and photoelectrode material applied in the degradation of various pollutants result from its high band gap energy, high refractive index, high optical transparency in the visible region, high oxygen storage capacity, and chemical reactivity. The other properties of CeO₂ which should be mentioned include its high thermal stability, high hardness, oxygen ion conductivity, special redox features, and easy conversion between Ce⁺³ and Ce⁺⁴ oxidation states [132-156].

The conduction band (CB) of g-C₃N₄ is more negative than that of CeO₂ (-1.24 eV and -0.44, respectively), while CeO₂ possesses a relatively positive valance band (VB) (2.56 eV) compared to the conduction band of g-C₃N₄, would theoretically facilitate the electron transition within the coupled photocatalyst to prolong the electron-hole separation [157]. Particularly, under the illumination of visible light, g-C₃N₄ can be excited to generate electron-hole pairs. Cerium (Ce) has exciting catalytic characteristics because 4d and 5p electrons sufficiently defend the 4f orbitals. The photogenerated electrons in the conduction band of CeO₂ tend to transfer and recombine with the photogenerated holes in the valence band of g-C₃N₄. Like this, the larger number of photogenerated electrons accumulated in the conduction band of g-C₃N₄ can reduce the adsorbed O₂ to form more O₂^{– ●}. At the same time, the photogenerated holes left behind in the valence band of CeO₂ can oxidize the adsorbed H₂O to give OH^●. But, the photocatalytic activity of the g-C₃N₄/CeO₂ system would be significantly increased, leading to the decomposition of organic compounds by O₂^{– ●} and OH^● reactive species.

In this study, a novel g-C₃N₄/CeO₂ NCs as a photocatalys was examined during photocatalytic degradation process in the efficient removal of OFX from pharmaceutical industry wastewater plant, İzmir, Turkey. Different pH values (3.0, 4.0, 6.0, 7.0, 9.0 and 11.0), increasing OFX concentrations (5 mg/l, 10 mg/l, 20 mg/l and 40 mg/l), increasing g-C₃N₄/CeCO₂ NCs concentrations (1 mg/l, 2 mg/l, 4 mg/l, 6 mg/l, 8 mg/l and 10 mg/l), different g-C₃N₄/CeO₂ NCs mass ratios (5/5, 6/4, 7/3, 8/2, 9/1, 1/9, 2/8, 3/7 and 4/6), increasing recycle times (1., 2., 3., 4., 5., 6. and 7.) was operated during photocatalytic degradation process in the efficient removal of OFX in pharmaceutical industry wastewater. The characteristics of the synthesized NPs were XRD, FESEM, EDX, FTIR, TEM and DRS analyses, respectively. The acute toxicity assays were operated with Microtox (Aliivibrio fischeri also called Vibrio fischeri) and Daphnia magna acute toxicity tests. The photocatalytic degradation mechanisms of g-C₃N₄/CeO₂ NCs and the reaction kinetics of OFX were evaluated in pharmaceutical industry wastewater during photocatalytic degradation process. ANOVA statistical analysis was used for all experimental samples.

Materıals and Methods

Characterization of Pharmaceutical Industry Wastewater

Characterization of the biological aerobic activated sludge proses from a pharmaceutical industry wastewater plant, İzmir, Turkey was performed. The results are given as the mean value of triplicate samplings (Table 1).

Table 1: Characterization of Pharmaceutical Industry Wastewater

Parameters	Unit	Concentrations
Chemical oxygen demand-total (CODtotal)	(mg/l)	4000
Chemical oxygen demand-dissolved (CODdissolved)	(mg/l)	3200
Biological oxygen demand-5 days (BOD5)	(mg/l)	1500
BOD5/CODdissolved		0.5
Total organic carbons (TOC)	(mg/l)	1800
Dissolved organic carbons (DOC)	(mg/l)	1100
pH		8.3
Salinity as Electrical conductivity (EC)	(mS/cm)	1552
Total alkalinity as CaCO3	(mg/l)	750
Total volatile acids (TVA)	(mg/l)	380
Turbidity (Nephelometric Turbidity unit, NTU)	NTU	7.2
Color	1/m	50
Total suspended solids (TSS)	(mg/l)	250
Volatile suspended solids (VSS)	(mg/l)	187
Total dissolved solids (TDS)	(mg/l)	825
Nitride (NO2–)	(mg/l)	1.7
Nitrate (NO3–)	(mg/l)	1.91
Ammonium (NH4+)	(mg/l)	2.3
Total Nitrogen (Total-N)	(mg/l)	3.2
SO3-2	(mg/l)	21.4
SO4-2	(mg/l)	29.3
Chloride (Cl–)	(mg/l)	37.4
Bicarbonate (HCO3–)	(mg/l)	161
Phosphate (PO4-3)	(mg/l)	16
Total Phosphorus (Total-P)	(mg/l)	40
Total Phenols	(mg/l)	70
Oil & Grease	(mg/l)	220
Cobalt (Co+3)	(mg/l)	0.2
Lead (Pb+2)	(mg/l)	0.4
Potassium (K+)	(mg/l)	17
Iron (Fe+2)	(mg/l)	0.42
Chromium (Cr+2)	(mg/l)	0.44
Mercury (Hg+2)	(mg/l)	0.35
Zinc (Zn+2)	(mg/l)	0.11

Preparation of Graphitic Carbon Nitride (g-C₃N₄) Nanoparticles

g-C₃N₄ was prepared by calcination of melamine (C₃H₆N₆) in a crucible with a lid at 550°C for 4 h. The obtained yellow powder was ground in an agate mortar after being cooled down to 25°C room temperature.

Preparation of Cerium Dioxide (CeO₂) Nanoparticles

CeO₂ NPs were prepared by sol–gel method. Nano-sized CeO₂ was also prepared by the sol–gel procedure using Cerium nitrate hexahydrate [Ce(NO₃)₃.6H₂O] and 20 ml of Triethanolamine (C₆H₁₅NO₃). Then, they were mixed together by a magnetic stirrer on a hot plate to insure that the cerium salt was dissolved in C₆H₁₅NO₃. After that the solution was heated up to 90°C until the clear dark brown homogenous solution, sol, was observed. To prepare black colloidal solution (gel), it was kept in a digital furnace at 270°C for 2 h. As gel was produced, it was cooled to 25°C room temperature. In order to form the expected precipitate, the volume of the gel solution was adjusted to 100 ml by adding ethanol (C₂H₆O). Then, synthesized precipitate was separated by centrifugation and washed by deionized water and C₂H₆O. Finally, the produced CeO₂ NPs was dried at 90°C and calcinated.

Preparation of A Novel Graphitic Carbon Nitride/Cerium Dioxide (g-C₃N₄/CeO₂) Nanocomposites

The g-C₃N₄/CeO₂ NCs was synthesized by the hydrothermal-calcination method. Firstly, 1 gram g-C₃N₄ NPs was added into distilled water and magnetically stirred for 30 min. Then, the portions of prepared g-C₃N₄ NPs were added to the mixtures to obtain the mass ratios of g-C₃N₄ to CeO₂ of 5/5, 6/4, 7/3, 8/2, 9/1, 1/9, 2/8, 3/7 and 4/6, respectively, and kept being stirred for another 1 h. The final mixtures were transferred into a 100 ml autoclave and reacted at 180°C for different hydrothermal (HT) times of 2 h, 4 h and 6 h. The final samples were centrifuged and washed with distilled water and C₂H₆O for 2 times. Then, the samples were dried, and finally, the dried products were heated in a Muffle furnace at different calcination temperatures of 300°C, 400°C and 500°C for 4 h to get the target composites. The synthesis conditions and the corresponding sample names were summarized at Table 2.

Table 2: The optimization parameters of g-C₃N₄/CeO₂ NCs samples

Sample Name	Mass Ratios of g-C3N4/CeO2 NCs	Calcination Temperature (°C) in 240 min	Hydrothermal Time (min) at 180°C
HT-2h-Cal300	8/2	300°C	120
HT-2h-Cal400	8/2	400°C	120
HT-2h-Cal500	8/2	500°C	120
HT-4h-Cal300	8/2	300°C	240
HT-4h-Cal400	8/2	400°C	240
HT-4h-Cal500	8/2	500°C	240
HT-6h-Cal300	8/2	300°C	360
HT-6h-Cal400	8/2	400°C	360
HT-6h-Cal500	8/2	500°C	360
5/5 wt, g-C3N4/CeO2	5/5	500°C	360
6/4 wt, g-C3N4/CeO2	6/4	500°C	360
7/3 wt, g-C3N4/CeO2	7/3	500°C	360
8/2 wt, g-C3N4/CeO2	8/2	500°C	360
9/1 wt, g-C3N4/CeO2	9/1	500°C	360
1/9 wt, g-C3N4/CeO2	1/9	500°C	360
2/8 wt, g-C3N4/CeO2	2/8	500°C	360
3/7 wt, g-C3N4/CeO2	3/7	500°C	360
4/6 wt, g-C3N4/CeO2	4/6	500°C	360

Photocatalytic Degradation Reactor

A 2 liter cylinder quartz glass reactor was used for the photodegradation experiments in the pharmaceutical industry wastewater at different operational conditions. 1000 ml pharmaceutical industry wastewater was filled for experimental studies and the photocatalyst were added to the cylinder quartz glass reactors. The UV-A lamps were placed to the outside of the photo-reactor with a distance of 3 mm. The photocatalytic reactor was operated with constant stirring (1.5 rpm) during the photocatalytic degradation process. 10 ml of the reacting solution were sampled and centrifugated (at 10000 rpm) at different time intervals. The UV irradiation treatments were created using one or three UV-A lamp emitting in the 350–400 nm range (λmax = 368 nm; FWHM = 17 nm; Actinic BL TL-D 18W, Philips). Six 50 W UV-A lamps (Total: 300 W UV-A lamps) were used during experimental conditions for this study.

Characterization

X-Ray Diffraction Analysis

Powder XRD patterns were recorded on a Shimadzu XRD-7000, Japan diffractometer using Cu Kα radiation (λ = 1.5418 Å, 40 kV, 40 mA) at a scanning speed of 1°/min in the 10-80° 2θ range. Raman spectrum was collected with a Horiba Jobin Yvon-Labram HR UV-Visible NIR (200-1600 nm) Raman microscope spectrometer, using a laser with the wavelength of 512 nm. The spectrum was collected from 10 scans at a resolution of 2 /cm. The zeta potential was measured with a SurPASS Electrokinetic Analyzer (Austria) with a clamping cell at 300 mbar.

Field Emission Scanning Electron Microscopy (FESEM) and Energy Dispersive X-Ray (EDX) Spectroscopy Analysis

The morphological features and structure of the synthesized catalyst were investigated by FESEM (FESEM, Hitachi S-4700), equipped with an EDX spectrometry device (TESCAN Co., Model III MIRA) to investigate the composition of the elements present in the synthesized catalyst.

Fourier Transform Infrared Spectroscopy (FTIR) Analysis

The FTIR spectra of samples was recorded using the FT-NIR spectroscope (RAYLEIGH, WQF-510).

Transmission Electron Microscopy (TEM) Analysis

The structure of the samples were analysed TEM analysis. TEM analysis was recorded in a JEOL JEM 2100F, Japan under 200 kV accelerating voltage. Samples were prepared by applying one drop of the suspended material in ethanol onto a carbon-coated copper TEM grid, and allowing them to dry at 25°C room temperature.

Diffuse Reflectance UV-Vis Spectra (DRS) Analysis

DRS Analysis in the range of 200–800 nm were recorded on a Cary 5000 UV-Vis Spectrophotometer from Varian. DRS was used to monitor the OFX antibiotic concentration in experimental samples.

Analytical Procedures

Chemical oxygen demand-total (COD_total), chemical oxygen demand-dissolved (COD_dissolved), total phosphorus (Total-P), phosphate phosphorus (PO₄^-3-P), total nitrogen (Total-N), ammonium nitrogen (NH₄⁺-N), nitrate nitrogen (NO₃^–-N), nitrite nitrogen (NO₂^–-N), biological oxygen demand 5-days (BOD₅), pH, Temperature [(°C)], total suspended solids (TSS), total volatile suspended solids (TVSS), total organic carbon (TOC), Oil, Chloride (Cl^–), total phenol, total volatile acids (TVA), disolved organic carbon (DOC), total alkalinity, turbidity, total dissolved solid (TDS), color, sulfide (SO₃^-2), sulfate (SO₄^-2), bicarbonate (HCO₃^–), salinity, cobalt (Co⁺³), lead (Pb⁺²), potassium (K⁺), iron (Fe⁺²), chromium (Cr⁺²), Mercury (Hg⁺²) and zinc (Zn⁺²) were measured according to the Standard Methods (2017) 5220B, 5220D, 4500-P, 4500-PO₄^-3, 4500-N, 4500-NH₄⁺, 4500-NO₃^–, 4500-NO₂^–, 5210B, 4500-H⁺, 2320, 2540D, 2540E, 5310, 5520, 4500-Cl^–, 5530, 5560B, 5310B, 2320, 2130, 2540E, 2120, 4500-SO₃^-2, 4500-SO₄^-2, 5320, 2520, 3500-Co⁺³, 3500-Pb⁺², 3500-K⁺, 3500-Fe⁺², 3500-Cr⁺², 3500-Hg⁺², 3500-Zn⁺², respectively [158].

Total-N, NH₄⁺-N, NO₃^–-N, NO₂^–-N, Total-P, PO₄^-3-P, total phenol, Co⁺³, Pb⁺², K⁺, Fe⁺², Cr⁺², Hg⁺², Zn⁺², SO₃^-2, and SO₄^-2 were measured with cell test spectroquant kits (Merck, Germany) at a spectroquant NOVA 60 (Merck, Germany) spectrophotometer (2003).

The measurement of color was carried out following the methods described by Olthof and Eckenfelder [159] and Eckenfelder [160]. According these methods, the color content was determined by measuring the absorbance at three wavelengths (445 nm, 540 nm and 660 nm), and taking the sum of the absorbances at these wavelengths. In order to identify the color in pharmaceutical industry wastewater (25 ml) was acidified at pH=2.0 with a few drops of 6 N HCl and extracted three times with 25 ml of ethyl acetate. The pooled organic phases were dehydrated on sodium sulphate, filtered and dried under vacuum. The residue was sylilated with bis(trimethylsylil)trifluoroacetamide (BSTFA) in dimethylformamide and analyzed by gas chromatography–mass spectrometry (GC-MS) and gas chromatograph (GC) (Agilent Technology model 6890N) equipped with a mass selective detector (Agilent 5973 inert MSD). Mass spectra were recorded using a VGTS 250 spectrometer equipped with a capillary SE 52 column (HP5-MS 30 m, 0.25 mm ID, 0.25 μm) at 220°C with an isothermal program for 10 min. The initial oven temperature was kept at 50°C for 1 min, then raised to 220°C at 25°C/min and from 200 to 300°C at 8°C/min, and was then maintained for 5.5 min. High purity He (g) was used as the carrier gas at constant flow mode (1.5 ml/min, 45 cm/s linear velocity).

The total phenol was monitored as follows: 40 ml of pharmaceutical industry wastewater was acidified to pH=2.0 by the addition of concentrated HCl. Total phenol was then extracted with ethyl acetate. The organic phase was concentrated at 40°C to about 1 ml and silylized by the addition of N,O-bis(trimethylsilyl) acetamide (BSA). The resulting trimethylsilyl derivatives were analysed by GC-MS (Hewlett-Packard 6980/HP5973MSD).

Methyl tertiary butyl ether (MTBE) was used to extract oil from the water and NPs. GC-MS analysis was performed on an Agilent gas chromatography (GC) system. Oil concentration was measured using a UV–vis spectroscopy fluorescence spectroscopy and a GC–MS (Hewlett-Packard 6980/HP5973MSD). UV–vis absorbance was measured on a UV–vis spectrophotometer and oil concentration was calculated using a calibration plot which was obtained with known oil concentration samples.

Acute Toxicity Assays

Microtox Acute Toxicity Test

Toxicity to the bioluminescent organism Aliivibrio fischeri (also called Vibrio fischeri or V. fischeri) was assayed using the Microtox measuring system according to DIN 38412L34, L341, (EPS 1/ RM/24 1992). Microtox testing was performed according to the standard procedure recommended by the manufacturer [161]. A specific strain of the marine bacterium, V. fischeri-Microtox LCK 491 kit was used for the Microtox acute toxicity assay. Dr. LANGE LUMIX-mini type luminometer was used for the microtox toxicity assay [162].

Daphnia magna Acute Toxicity Test

To test toxicity, 24-h born Daphnia magna were used as described in Standard Methods sections 8711A, 8711B, 8711C, 8711D and 8711E, respectively [163]. After preparing the test solution, experiments were carried out using 5 or 10 Daphnia magna introduced into the test vessels. These vessels had 100 ml of effective volume at 7.0– 8.0 pH, providing a minimum dissolved oxygen (DO) concentration of 6 mg/l at an ambient temperature of 20–25°C. Young Daphnia magna were used in the test (≤24 h old); 24–48 h exposure is generally accepted as standard for a Daphnia magna acute toxicity test. The results were expressed as mortality percentage of the Daphnia magna. Immobile animals were reported as dead Daphnia magna.

Statistical Analysis

ANOVA analysis of variance between experimental data was performed to detect F and P values. The ANOVA test was used to test the differences between dependent and independent groups, [164]. Comparison between the actual variation of the experimental data averages and standard deviation is expressed in terms of F ratio. F is equal (found variation of the date averages/expected variation of the date averages). P reports the significance level, and d.f indicates the number of degrees of freedom. Regression analysis was applied to the experimental data in order to determine the regression coefficient R², [165]. The aforementioned test was performed using Microsoft Excel Program.

All experiments were carried out three times and the results are given as the means of triplicate samplings. The data relevant to the individual pollutant parameters are given as the mean with standard deviation (SD) values.

Results and Dıscussıons

A Novel g-C₃N₄/CeO₂ NCs Characteristics

The Results of X-Ray Diffraction (XRD) Analysis

The results of XRD analysis was observed to pure g-C₃N₄ NPs, pure CeO₂ NPs and g-C₃N₄/CeO₂ NCs, respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 1). The characterization peaks were observed at 2θ values of 14.21°, 20.12° and 28.24°, respectively, implying pure g-C₃N₄ NPs in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 1a). The characterization peaks were obtained at 2θ values of 29.41°, 34.22°, 48.45°, 57.62°, 59.27°, 70.18°, 77.17° and 79.31°, respectively, implying pure CeO₂ NPs in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 1b). The characterization peaks were found at 2θ values of 13.20°, 28.72°, 33.67°, 48.15°, 58.39°, 60.16°, 71.17°, 75.35° and 79.53°, respectively, and which can also be indexed as (100), (002), (111), (200), (220), (311), (222), (400), (331) and (420), respectively, implying g-C₃N₄/CeO₂ NCs in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 1c).

Figure 1: The XRD patterns of (a) pure g-C₃N₄ NPs (black pattern), (b) pure CeO₂ NPs (blue pattern) and (c) g-C₃N₄/CeO₂ NCs (red pattern), respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

The Results of Diffuse Reflectance UV-Vis Spectra (DRS) Analysis

The absorption spectra of OFX was observed in DRS Analysis (Figure 2). First, the absorption spectra of OFX were obtained at a maximum concentration of 40 mg/l in the wavelength range from 250 nm to 800 nm using diffuse reflectance UV-Vis spectra (Figure 2). Absorption peaks were observed at wavelengths of 400 nm for pure g-C₃N₄ NPs (black pattern) (Figure 2a), 310 nm for pure CeO₂ NPs (green pattern) (Figure 2b), and 340 nm for g-C₃N₄/CoMoO₄ NCs (blue pattern) (Figure 2c), respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

Figure 2: The DRS patterns of (a) pure g-C₃N₄ NPs (black pattern) (b) pure CeO₂ NPs (green pattern) and (c) g-C₃N₄/CeO₂ NCs (blue pattern), respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

The Results of Field Emission Scanning Electron Microscopy (FESEM) Analysis

The morphological features of pure g-C₃N₄ NPs, pure CeO₂ NPs and g-C₃N₄/CeO₂ NCs were characterized through FE-SEM images (Figure 3). The FESEM images of pure g-C₃N₄ NPs were obtained in in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 3a). The FESEM images of pure CeO₂ NPs were observed in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 3b). The FESEM images of g-C₃N₄/CeO₂ NCs were characterized in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 3c).

Figure 3: FESEM images of (a) pure g-C₃N₄ NPs, (b) pure CeO₂ NPs and (c) g-C₃N₄/CeO₂ NCs, respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

The Results of Energy Dispersive X-Ray (EDX) Spectroscopy Analysis

The EDX analysis was also performed to investigate the composition of g-C₃N₄/CeO₂ NCs (Figure 4), respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

Figure 4: EDX spectrum of g-C₃N₄/CeO₂ NCs, respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

The Results of Fourier Transform Infrared Spectroscopy (FTIR) Analysis

The FTIR spectrum of pure g-C₃N₄ NPs (black spectrum), pure CeO₂ NPs (blue spectrum) and g-C₃N₄/CeO₂ NCs (red spectrum), respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 5). The main peaks of FTIR spectrum for pure g-C₃N₄ NPs (black spectrum) was observed at 1645 1/cm, 1564 1/cm, 1411 1/cm, 1321 1/cm, 1240 1/cm and 807 1/cm wavenumber, respectively (Figure 5a). The main peaks of FTIR spectrum for pure CeO₂ NPs (blue spectrum) was obtained at 462 1/cm wavenumber, respectively (Figure 5b). The main peaks of FTIR spectrum for g-C₃N₄/CeO₂ NCs (red spectrum) was determined at 462 1/cm wavenumber, respectively (Figure 5c).

Figure 5: FTIR spectrum of (a) pure g-C₃N₄ NPs (black spectrum), (b) pure CeO₂ NPs (blue spectrum) and (c) g-C₃N₄/CeO₂ NCs (red spectrum), respectively, in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

The Results of Transmission Electron Microscopy (TEM) Analysis

The TEM images of g-C₃N₄/CeO₂ NCs was observed in micromorphological structure level in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal (Figure 6).

Figure 6: TEM images of g-C₃N₄/CeO₂ NCs in micromorphological structure level in pharmaceutical industry wastewater with photocatalytic degradation process for OFX antibiotic removal.

The Reaction Kinetics of OFX Antibiotic

The reaction kinetics OFX were investigated using the Langmuir–Hinshelwood first-order kinetic model, expressed by Eddy et al. [166], as following Equation (1):

where; r_o: denotes the initial photocatalytic degradation reaction rate (mg/l.min), and k: denotes the rate constant of a first-order reaction. At the beginning of the reaction, t = 0, C_t = C₀, the equation can be obtained after integration as following Equation (2):

where; C₀ and C : are the initial and final concentration (mg/l) of OFX; the solution at t (min) and k (1/min) are the rate constant.

The correlation coefficients had R² values greater than 0.9, as a result, the first-order kinetic model fit the experimental data well. The first-order rate constants (k) were determined from the slope of the linear plots.

Photocatalytic Degradation Mechanisms

The photocatalytic performance of the catalyst in the degradation of OFX is determined by photons. The degradation mechanism of OFX by hydroxyl radicals (OH^●) radicals concerning g-C₃N₄/CeO₂ NCs as following equations (Equation 3, Equation 4, Equation 5, Equation 6, Equation 7, Equation 8, Equation 9 and Equation 10):

g-C₃N₄/CeO₂ NCs absorbs photons with energies greater than the photocatalyst bandgap. As a result, the electron in the valence band (VB) jumps to the conduction band (CB), leaving a hole in the CB. The electrons present in the CB and VB will react with oxygen (O₂) and water (H₂O) molecules which are absorbed by the photocatalyst and lead to the formation of OH^● radicals which react with OFX. OH^● radicals are produced when the photocatalyst surface is illuminated with photons, and OH^● radicals are strong oxidising species, with an oxidation potential of approximately 2.8 V [as opposed to Normal hydrogen electrode (NHE)], which may increase total pollutant mineralisation. Normally, the higher the rate of formation of OH^● radicals, the greater the separation efficiency of electron-hole pairs. In this way, there is a correlation between the increased photocatalytic activity and the rate of formation of OH^● radicals. The OH^● radicals generation of g-C₃N₄/CeO₂ NCs was extremely high, indicating that the sample has a high electron and hole separation rate.

CeO₂ composites with g-C₃N₄ are also promising photocatalytic materials with a lower band gap energy [167-169] and significantly higher photocatalytic efficiency in degradation processes [170,171]. Considering the position of CB and VB in CeO₂ and g-C₃N₄, the higher photocatalytic efficiency can be attributed to the transfer of photoexcited electrons and holes between CeO₂ and g-C₃N₄, which suppresses the recombination of photogenerated h⁺/e ⁻ pairs. During irradiation, photogenerated electrons on CB in g-C₃N₄ are transferred to CB in CeO₂ and react with O₂, while photogenerated holes on VB in CeO₂ are transferred to VB in g-C₃N₄ and react with H₂O according to the following reactions [172]:

The superoxide and hydroxyl radicals formed in the above-presented reactions take part in the degradation of pollutants. In the case of CeO₂ composites with g-C₃N₄, two problems have still not been resolved. The first one is related to the lower rates of TOC or COD decrease in wastewater in comparison with the degradation rate of pollutants [173]. The second one is attributed to the immobilization of a composite photocatalyst, which could eliminate the post-treatment process of photocatalyst removal from the wastewater.

Effect of Increasing pH values for OFX Removal in Pharmaceutical Industry Wastewater during Photocatalytic Degradation Process

Increasing pH values (pH=3.0, pH=4.0, pH=6.0, pH=7.0, pH=9.0 and pH=11.0, respectively) was examined during photocatalytic degradation process in pharmaceutical industry wastewater for OFX removal (Figure 7). 67.2%, 85.7%, 96.4%, 56.5% and 44.8% OFX removal efficiencies was measured at pH=3.0, pH=4.0, pH=6.0, pH=7.0 and pH=11.0, respectively, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at 25°C (Figure 7). The maximum 99% OFX removal efficiency was obtained during photocatalytic degradation process in pharmaceutical industry wastewater, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 7).

Figure 7: Effect of increasing pH values for OFX removal in pharmaceutical industry wastewater during photocatalytic degradation process, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

Effect of Increasing OFX Concentrations for OFX Removal in Pharmaceutical Industry Wastewater during Photocatalytic Degradation Process

Increasing OFX concentrations (5 mg/l, 10 mg/l, 20 mg/l and 40 mg/l) were operated at 300 W UV-vis irradiation power, after 180 min photocatalytic degradation time, at pH=6.0, at 25°C, respectively (Figure 8). 85.3%, 94.1% and 77.2% OFX removal efficiencies were obtained to 5 mg/l, 10 mg/l and 40 mg/l OFX concentrations, respectively, at pH=6.0 and at 25°C (Figure 8). The maximum 99% OFX removal efficieny was found with photocatalytic degradation process in pharmaceutical industry wastewater, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 8).

The percentage decrease (8%) in the concentration of OFX during the studies under the dark conditions was due to the contaminant adsorption onto the catalyst surface [174]. The formation of contaminant monolayer on the surface of the catalyst may have occupied all its active sites, and therefore no more adsorption was observed.

Figure 8: Effect of increasing OFX concentrations for OFX removal in pharmaceutical industry wastewater during photocatalytic degradation process, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

Effect of Increasing g-C₃N₄/CeO₂ NCs Concentrations for OFX Removals in Pharmaceutical Industry Wastewater during Photocatalytic Degradation Process

Increasing g-C₃N₄/CeO₂ NCs concentrations (1 mg/l, 2 mg/l, 4 mg/l, 6 mg/l, 8 mg/l and 10 mg/l) were operated at 20 mg/l OFX, at 150 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0, at 25°C, respectively (Figure 9). 54.5%, 68.1%, 75.8%, 87.3% and 92.1% OFX removal efficiencies were obtained to 1 mg/l, 2 mg/l, 4 mg/l, 6 mg/l and 10 mg/l g-C₃N₄/CeO₂ NCs concentrations, respectively, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0, at 25°C, respectively (Figure 9). The maximum 99% OFX removal efficieny was measured to 8 mg/l g-C₃N₄/CeO₂ NCs with photocatalytic degradation process in pharmaceutical industry wastewater, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 9).

Figure 9: Effect of increasing g-C₃N₄/CeO₂ NCs concentrations for OFX removal in pharmaceutical industry wastewater during photocatalytic degradation process, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

Effect of Different g-C₃N₄/CeO₂ NCs Mass Ratios for OFX Removals in Pharmaceutical Industry Wastewater during Photocatalytic Degradation Process

Different g-C₃N₄/CeO₂ mass ratios (5/5wt, 6/4wt, 7/3wt, 8/2wt, 9/1wt, 1/9wt, 2/8wt, 3/7wt and 4/6wt, respectively) were examined for OFX removal in pharmaceutical industry wastewater during photocatalytic degradation process, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 10). 80.3%, 84.6%, 77.9%, 62.1%, 48.4%, 55.2%, 64.0% and 79.7% OFX removal efficiencies were measured at 5/5wt, 6/4 wt, 7/3wt, 8/2wt, 9/1wt, 1/9wt, 3/7wt and 4/6wt g-C₃N₄/CeO₂ NCs mass ratios, respectively, at 20 mg/l OFX after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 10). The maximum 99% OFX removal efficiency was measured at 2/8wt g-C₃N₄/CeO₂ NCs mass ratios at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 10).

Figure 10: Effect of different g-C₃N₄/CeO₂ NCs mass ratios for OFX removal in pharmaceutical industry wastewater during photocatalytic degradation process, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

Effect of Different Recycle Times for OFX Removals in Pharmaceutical Industry Wastewater during Photocatalytic Degradation Process

Different recycle times (1., 2., 3., 4., 5., 6. and 7.) were operated for OFX removals in pharmaceutical industry wastewater during photocatalytic degradation process, at 20 mg/l OFX, 8 mg/l g-C₃N₄/CeO₂ NCs, at 2/8wt g-C₃N₄/CeO₂ NCs mass ratio, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 11). 97.5%, 96.2%, 94%, 93.8%, 89.2%, 86.2% and 80.1% OFX removal efficiencies were measured after 2. recycle time, 3. recycle time, 4. recycle time, 5. recycle time, 6. recycle time and 7. recycle time, respectively, at 20 mg/l OFX, 8 mg/l g-C₃N₄/CeO₂ NCs, at 2/8wt g-C₃N₄/CeO₂ NCs mass ratio, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 11). The maximum 99% OFX removal efficiency was measured in pharmaceutical industry wastewater during photocatalytic degradation process, after 1. recycle time, at 20 mg/l OFX, 8 mg/l g-C₃N₄/CeO₂ NCs, at 2/8wt g-C₃N₄/CeO₂ NCs mass ratio, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively (Figure 11).

Figure 11: Effect of recycle times for OFX removal in pharmaceutical industry wastewater during photocatalytic degradation process, at 20 mg/l OFX, 8 mg/l g-C₃N₄/CeO₂ NCs, at 2/8wt g-C₃N₄/CeO₂ NCs mass ratio, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

Acute Toxicity Assays

Effect of Increasing OFX Concentrations on the Microtox (Aliivibrio fischeri or Vibrio fischeri) Acute Toxicity Removal Efficiencies in Pharmaceutical Industry Wastewater at Increasing Photocatalytic Degradation Time and Temperature

In Microtox with Aliivibrio fischeri (also called Vibrio fischeri) acute toxicity test, the initial EC₉₀ values at pH=7.0 was found as 825 mg/l at 25°C (Table 3: SET 1). After 60 min, 120 min and 180 min photocatalytic degradation time, the EC₉₀ values decreased to EC₅₇=414 mg/l to EC₂₂=236 mg/l and to EC₁₂=165 mg/l in OFX=20 mg/l at 30°C (Table 3: SET 3). The Microtox (Aliivibrio fischeri) acute toxicity removal efficiencies were 40.86%, 79.75% and 90.86% after 60 min, 120 min and 180 min, respectively, in OFX=20 mg/l and at 30°C (Table 3: SET 3).

The EC₉₀ values decreased to EC₅₁, to EC₁₆ and to EC₆ after 60 min, 120 min and 180 min, respectively, in OFX=20 mg/l, at 60°C (Table 3: SET 3). The EC₅₁, the EC₁₁ and the EC₇ values were measured as 550 mg/l, 540 mg/l and 500 mg/l, respectively, in OFX=20 mg/l at 60°C. The toxicity removal efficiencies were 46.41%, 85.30% and 96.41% after 60 min, 120 min and 180 min, respectively, in OFX=20 mg/l, at 60°C (Table 3: SET 3). 96.41% maximum Microtox (Aliivibrio fischeri) acute toxicity removal yield was found in OFX=20 mg/l after 180 min and at 60°C (Table 3: SET 3).

The EC₉₀ values decreased to EC₆₂=422 mg/l to EC₂₁=241 mg/l and to EC₁₇=168 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=5 mg/l at 30°C (Table 3: SET 3). The EC₉₀ values decreased to EC₆₂=421 mg/l to EC₂₇=239 mg/l and to EC₁₁=167 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=10 mg/l at 30°C. The EC₉₀ values decreased to EC₆₇=408 mg/l to EC₃₂=230 mg/l and to EC₂₂=162 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=40 mg/l at 30°C. The Microtox (Aliivibrio fischeri or Vibrio fischeri) acute toxicity removals were 85.30%, 85.28% and 79.75% in 5 mg/l, 10 mg/l and 40 mg/l OFX, respectively, after 180 min, at 30°C. It was obtained an inhibition effect of OFX=40 mg/l to Vibrio fischeri after 180 min and at 30°C (Table 3: SET 3).

The EC₉₀ values decreased to EC₅₇=419 mg/l to EC₂₂=266 mg/l and to EC₁₂=150 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=5 mg/l at 60°C (Table 3: SET 3). The EC₉₀ values decreased to EC₅₇=414 mg/l to EC₂₂=232 mg/l and to EC₁₂=161 mg/l after 60 min, 120 and 180 min, respectively, in OFX=10 mg/l at 60°C. The EC₉₀ values decreased to EC₆₂=403 mg/l to EC₂₇=218 mg/l and to EC₁₇=148 mg/l after 60 min, 120 and 180 min, respectively, in OFX=40 mg/l at 60°C. The Microtox (Aliivibrio fischeri or Vibrio fischeri) acute toxicity removals were 90.86%, 90.83% and 85.30% in 5 mg/l, 10 mg/l and 40 mg/l OFX, respectively, after 180 min, at 60°C. It was observed an inhibition effect of OFX=40 mg/l to Microtox with Vibrio fischeri after 180 min, and at 60°C (Table 3: SET 3).

Table 3: Effect of increasing OFX concentrations on Microtox (Aliivibrio fischeri) acute toxicity in pharmaceutical industry wastewater after photocatalytic degradation process, at 30°C and at 60°C, respectively.

No	Parameters	Microtox (Aliivibrio fischeri) Acute Toxicity Values, * EC (mg/l)
		25°C
		0 min		60 min		120 min		180 min
		*EC90		*EC		*EC		*EC
1	Raw ww, Control	825		EC70=510		EC60=650		EC49=638
		30°C				60°C
		0. min	60 min	120. min	180. min	0 min	60 min	120 min	180 min
		*EC90	*EC	*EC	*EC	*EC90	*EC	*EC	*EC
2	Raw ww, control	825	EC70=580	EC50=580	EC39=548	825	EC55=550	EC40=590	EC29=688
3	OFX=5 mg/l	825	EC62=422	EC27=242	EC17=168	825	EC57=419	EC22=266	EC12=150
	OFX=10 mg/l	825	EC62=421	EC27=239	EC17=167	825	EC57=414	EC22=232	EC12=161
	OFX=20 mg/l	825	EC57=414	EC22=236	EC12=165	825	EC52=550	EC17=540	EC7=500
	OFX=40 mg/l	825	EC67=408	EC32=230	EC22=162	825	EC62=403	EC27=218	EC17=148
* EC values were calculated based on CODdis (mg/l).

Effect of Increasing OFX Concentrations on the Daphnia magna Acute Toxicity Removal Efficiencies in Pharmaceutical Industry Wastewater at Increasing Photocatalytic Degradation Time and Temperature

The initial EC₅₀ values were observed as 850 mg/l at 25°C (Table 4: SET 1). After 60 min, 120 and 180 min photocatalytic degradation time, the EC₅₀ values decreased to EC₃₁=350 mg/l to EC₁₇=240 mg/l and to EC₁₂=90 mg/l in OFX=20 mg/l, at 30°C (Table 4: SET 3). The toxicity removal efficiencies were 42.96%, 72.87% and 82.65% after 60 min, 120 min and 180 min, respectively, in OFX=20 mg/l at 30°C (Table 4: SET 3).

The EC₅₀ values decreased to EC₂₇ to EC₁₂ and to EC₇ after 60 min, 120 min and 180 min, respectively, in OFX=20 mg/l at 60°C (Table 4: SET 3). The EC₂₇, the EC₁₂ and the EC₇ values were measured as 150 mg/l, 60 mg/l and 375 mg/l, respectively, in OFX=20 mg/l at 60°C. The toxicity removal efficiencies were 52.94%, 82.62% and 92.36% after 60 min, 120 min and 180 min, respectively, in OFX=20 mg/l at 60°C (Table 4: SET 3). 92.38% maximum Daphnia magna acute toxicity removal was obtained in OFX=20 mg/l after 180 min and at 60°C, respectively (Table 4: SET 3).

The EC₅₀ values decreased to EC₃₇=450 mg/l to EC₂₂=145 mg/l and to EC₁₇=260 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=5 mg/l at 30°C (Table 4: SET 3). The EC₅₀ values decreased to EC₃₇=450 mg/l to EC₂₂=175 mg/l and to EC₁₇=100 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=10 mg/l and at 30°C. The EC₅₀ values decreased to EC₄₂=300 mg/l to EC₂₇=170 mg/l and to EC₂₂=52 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=40 mg/l and at 30°C. The Daphnia magna acute toxicity removals were 72.22%, 72.56% and 63.21% in 5 mg/l, 10 mg/l and 40 mg/l OFX, respectively, after 180 min and at 30°C. It was observed an inhibition effect of OFX=40 mg/l to Daphnia magna after 180 min and at 30°C (Table 4: SET 3).

The EC₅₀ values decreased to EC₃₂=130 mg/l to EC₁₇=425 mg/l and to EC₁₂=340 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=5 mg/l and at 60°C (Table 4: SET 3). The EC₅₀ values decreased to EC₃₂=425 mg/l to EC₁₇=140 mg/l and to EC₇=90 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=10 mg/l and at 60°C. The EC₅₀ values decreased to EC₃₇=250 mg/l to EC₂₂=110 mg/l and to EC₁₇=10 mg/l after 60 min, 120 min and 180 min, respectively, in OFX=40 mg/l and at 60°C. The Daphnia magna acute toxicity removals were 83.06%, 92.65% and 73.11% in 5 mg/l, 10 mg/l and 40 mg/l OFX, respectively, after 180 min and at 60°C. It was observed an inhibition effect of OFX=40 mg/l to Daphnia magna after 180 min and at 60°C (Table 4: SET 3).

Increasing the OFX concentrations from 5 mg/l to 40 mg/l did not have a positive effect on the decrease of EC₅₀ values as shown in Table 4 at SET 3. OFX concentrations > 20 mg/l decreased the acute toxicity removals by hindering the photocatalytic degradation process. Similarly, a significant contribution of increasing OFX concentration to acute toxicity removal at 60°C after 180 min of photocatalytic degradation time was not observed. Low toxicity removals found at high OFX concentrations could be attributed to their detrimental effect on the Daphnia magna (Table 4: SET 3).

Table 4: Effect of increasing OFX concentrations on Daphnia magna acute toxicity in pharmaceutical industry wastewater after photocatalytic degradation process, at 30°C and at 60°C.

No	Parameters	Daphnia magna Acute Toxicity Values, * EC (mg/l)
		25°C
		0. min			60. min		120. min		180. min
		*EC50			*EC		*EC		*EC
1	Raw ww, control	850			EC45=625		EC40=370		EC29=153
		30°C					60°C
		0 min	60 min	120 min		180. min	0. min	60. min	120. min	180. min
		*EC50	*EC	*EC		*EC	*EC50	*EC	*EC	*EC
2	Raw ww, control	850	EC39=468	EC34=228		EC23=111	850	EC34=373	EC29=210	EC18=71
3	OFX=5 mg/l	850	EC32=450	EC22=145		EC17=260	850	EC32=130	EC17=425	EC12=340
	OFX=10 mg/l	850	EC37=450	EC22=175		EC17=100	850	EC32=425	EC17=140	EC7=90
	OFX=20 mg/l	850	EC32=350	EC17=240		EC12=90	850	EC27=150	EC12=60	EC7=375
	OFX=40 mg/l	850	EC42=300	EC27=170		EC22=52	850	EC37=250	EC22=110	EC17=11
* EC values were calculated based on CODdis (mg/l).

Direct Effects of OFX Concentrations on the Acute Toxicity of Microtox (Aliivibrio fischeri or Vibrio fischeri) and Daphnia magna without Pharmaceutical Industry Wastewater after Photocatalytic Degradation Process

The acute toxicity test was performed in the samples containing 5 mg/l, 10 mg/l, 20 mg/l and 40 mg/l OFX concentrations, at 25°C room temperature. In order to detect the direct responses of Microtox (Aliivibrio fischeri or Vibrio fischeri) and Daphnia magna to the increasing OFX concentrations the toxicity test were performed without pharmaceutical industry wastewater after photocatalytic degradation process, at 25°C room temperature. The initial EC values and the the EC₅₀ values were measured in the samples containing increasing OFX concentrations after 180 min photocatalytic degradation time. Table 5 showed the responses of Microtox (Aliivibrio fischeri or Vibrio fischeri) and Daphnia magna to increasing OFX concentrations.

The acute toxicity originating only from 5 mg/l, 10 mg/l, 20 mg/l and 40 mg/l OFX were found to be low (Table 5). 5 mg/l OFX did not exhibited toxicity to Aliivibrio fischeri (or Vibrio fischeri) and Daphnia magna before and after 180 min photocatalytic degradation time. The toxicity atributed to the 10 mg/l, 20 mg/l and 40 mg/l OFX were found to be low in the samples without pharmaceutical industry wastewater after photocatalytic degradation process for the test organisms mentioned above. The acute toxicity originated from the OFX decreased significantly to EC₂, EC₄ and EC₆ after 180 min photocatalytic degradation time. Therefore, it can be concluded that the toxicity originating from the OFX is not significant and the real acute toxicity throughout photocatalytic degradation process was attributed to the pharmaceutical industry wastewater, to their metabolites and to the photocatalytic degradation by-products (Table 5).

Table 5: The responses of Microtox (Aliivibrio fischeri or Vibrio fischeri) and Daphnia magna acute toxicity tests in addition of increasing OFX concentrations without phamaceutical industry wastewater during photocatalytic degradation process after 180 min photocatalytic degradation time, at 25°C room temperature.

OFX Conc. (mg/l)	Microtox (Aliivibrio fischeri or Vibrio fischeri) Acute Toxicity Test			Daphnia magna Acute Toxicity Test
	Initial Acute Toxicity EC50 Value (mg/l)	Inhibitions after 180 min photocatalytic degradation time	EC Values (mg/l)	Initial Acute Toxicity EC50 Value (mg/l)	Inhibitions after 180 min photocatalytic degradation time	EC Values (mg/l)
5	EC10=24	–	–	EC10=39	–	–
10	EC15=79	3	EC2=3	EC20=99	5	EC3=5
20	EC20=149	5	EC4=6	EC30=199	6	EC6=11
40	EC25=219	7	EC6=9	EC40=299	9	EC8=15

The Comparison with Other Scientific Studies in the Literature

Comparison of our study “The use of a novel graphitic carbon nitride/cerium dioxide (g-C₃N₄/CeO₂) nanocomposites for the ofloxacin removal by photocatalytic degradation in pharmaceutical industry wastewaters and the evaluation of microtox (Aliivibrio fischeri) and Daphnia magna acute toxicity assays” with other scientific studies in the literature is summaried at Table 6 [175-182].

Table 6: The Comparison with other Scientific Studies in the Literature

Photocatalyst	Experimental Conditions (for maximum removal efficiencies)	Experimental Results	References
a-Bi2O3/g-C3N4	[DOX]=10 mg/l, [Material]=500 mg/l, [H2O2]=10 mM, Unadjusted pH, Xe lamp (150 W).	79.1% DOX (30 min)	(Liu et al., 2021a)
Ag/AgCl@ZIF-8/g-C3N4	150 W, Xe, λ > 420 nm, 50 mg/l, [LVFX]=10 mg/l, V=50 ml,	87.3% LVFX (60 min)	(Zhou et al., 2019)
Ag@ZIF-8/g-C3N4	300 W, Xe, λ > 420 nm, [Ten antibiotics] =10 mg/l, V=50 ml	90% (60 min)	(Guo et al., 2022)
Peroxymonosulfate/ZnFe2O4	Waters e2695 HPLC instrument (Milford, USA), UV-Vis detector λ=294 nm [OFX]=1000 mg/l, pH=6.0	80.9% OFX (30 min), pH 6.0	(Sun et al., 2021b)
Bi2WO6 and g-C3N4 nanosheets	[CRO]=16.5-66 µM, KrCl excilamp, λ=222 nm, 23 W, incident irradiance 0.74 mW/cm2, 60 min	91% Ceftriaxone (60 min)	(Sizykh et al., 2023)
Bi2WO6/g-C3N4	[CRO]=100 mg/l 300 W Xe lamp,	94.5% Ceftriaxone (120 min)	(Zhao et al., 2018)
CeO2-ZnO hetero photocatalyst	[TCN]=100 mg/l, 300 W Xenon lamp,	87.25% Tetracycline (60 min)	(Ye et al., 2016)
g-C3N4/CeO2 core-shell structure	Hydrothermal method, [DOX]=1000 mg/l, HCl=10 mg/l, H2O2=100 µl, 150 W Xe lamp (λ > 400 nm), g-C3N4=2.82 eV, CeO2=2.76 eV	66.7%g-C3N4, 71.7%CeO2, 84% g-C3N4/CeO2 (60 min)	(Liu et al., 2019)
CeO2/ATP/g-C3N4 ATP—attapulgite	Electrostatic-induced self-assembly method, Dibenzothiophene (DBT), m(catal)/m(DBT)=1/10, SO₂=200 mg/l, 30% H2O2, 300 W Xe lamp (λ > 420 nm)	Desulfurization 42% g-C3N4, 83% CeO2/g-C3N4, 98% CeO2/ATP/g-C3N4 (180 min)	(Li et al., 2017b)
g-C3N4/CeO2 NCs	g-C3N4/CeO2 NCs was prepared to hydrothermal calcination method, CeO2 was prepared sol-gel method, g-C3N4 was prepared to calcination method, pH=6.0, [OFX]=20 mg/l, [g-C3N4/CeO2 NCs]=8 mg/l, g-C3N4/CeO2 mass ratio=2/8, 300 W UV-vis A lamp λ=350-400 nm range (λmax=368 nm; FWHM=17 nm; Actinic BL TL-D 18W, Philips) Recycle time=7, at 25°C Microtox (Aliivibrio fischeri) and Daphnia magna acute toxicity assays	99% OFX (180 min, at 25°C) 96.41% maximum Microtox (Aliivibrio fischeri) acute toxicity removal (180 min, at 60°C) 92.38% maximum Daphnia magna acute toxicity removal (180 min, at 60°C), 99% OFX after 1. recycle time	This study
DOX: doxycycline; LVFX: Levofloxacin; CRO: ceftriaxone; TCN: Tetracycline; OFX: ofloxacin

Conclusıons

The maximum 99% OFX removal efficiency was obtained during photocatalytic degradation process in pharmaceutical industry wastewater, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

The maximum 99% OFX removal efficieny was found with photocatalytic degradation process in pharmaceutical industry wastewater, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

The maximum 99% OFX removal efficieny was measured to 8 mg/l g-C₃N₄/CeO₂ NCs with photocatalytic degradation process in pharmaceutical industry wastewater, at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

The maximum 99% OFX removal efficiency was measured at 2/8wt g-C₃N₄/CeO₂ NCs mass ratios at 20 mg/l OFX, at 300 W UV-vis light irradiation power, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

The maximum 99% OFX removal efficiency was measured in pharmaceutical industry wastewater during photocatalytic degradation process, after 1. recycle time, at 20 mg/l OFX, 8 mg/l g-C₃N₄/CeO₂ NCs, at 2/8wt g-C₃N₄/CeO₂ NCs mass ratio, after 180 min photocatalytic degradation time, at pH=6.0 and at 25°C, respectively.

96.41% maximum Microtox (Aliivibrio fischeri) acute toxicity removal yield was found in OFX=20 mg/l after 180 min and at 60°C. It was observed an inhibition effect of OFX=40 mg/l to Microtox with Vibrio fischeri after 180 min photocatalytic degradation time and at 60°C. 92.38% maximum Daphnia magna acute toxicity removal was obtained in OFX=20 mg/l after 180 min photocatalytic degradation time and at 60°C, respectively. It was observed an inhibition effect of OFX=40 mg/l to Daphnia magna after 180 min photocatalytic degradation time and at 60°C. OFX concentrations > 20 mg/l decreased the acute toxicity removals by hindering the photocatalytic degradation process. Similarly, a significant contribution of increasing OFX concentrations to acute toxicity removal at 60°C after 180 min photocatalytic degradation time was not observed. Finally, it can be concluded that the toxicity originating from the OFX is not significant and the real acute toxicity throughout photocatalytic degradation process was attributed to the pharmaceutical industry wastewater, to their metabolites and to the photocatalytic degradation process by-products.

As a result, the a novel g-C₃N₄/CeO₂ NCs photocatalyst during photocatalytic degradation process in pharmaceutical industry wastewater was stable in harsh environments such as acidic, alkaline, saline, and then was still effective process. When the amount of contaminant was increased, the a novel g-C₃N₄/CeO₂ NCs photocatalys during photocatalytic degradation process performance was still considerable. The synthesis and optimization of g-C₃N₄/CeO₂ heterostructure photocatalyst provides insights into the effects of preparation conditions on the material’s characteristics and performance, as well as the application of the effectively designed photocatalyst in the removal of antibiotics, which can potentially be deployed for purifying wastewater, especially pharmaceutical wastewater. Finally, the combination of a simple, easy operation preparation process, excellent performance and cost effective, makes this a novel g-C₃N₄/CeO₂ NCs a promising option during photocatalytic degradation process in pharmaceutical industry wastewater treatment.

Acknowledgement

This research study was undertaken in the Environmental Microbiology Laboratories at Dokuz Eylül University Engineering Faculty Environmental Engineering Department, Izmir, Turkey. The authors would like to thank this body for providing financial support.

References

1. Idham MF, Abdullah B, Yusof KM (2017) Effects of two cycle heat treatment on the microstructure and hardness of ductile iron. Pertanika J Sci Technol 25: 99-106.
2. Idham MF, Falyouna O, Eljamal O (2021) Effect of graphene oxide synthesis method on the adsorption performance of pharmaceutical contaminants. Proc Int Exch Innov Conf Eng Sci 7: 232-239.
3. Arenas NE, Melo VM (2018) Producción pecuaria y emergencia de antibiótico resistencia en Colombia: Revisión sistemática. Infectio 22: 110-119.
4. Pellerito A, Ameen SM, Micali M, Caruso G (2018) Antimicrobial substances for food packaging products: the current situation. J AOAC Int 101: 942-947. [crossref]
5. Fridkin S, Baggs, J, Fagan R, Magill S, Pollack LA, et al. (2014) Vital signs: improving antibiotic use among hospitalized patients. MMWR-Morb Mortal Wkly Rep 63: 194-200. [crossref]
6. Tamma PD, Avdic E, Li DX, Dzintars K, Cosgrove SE (2017) Association of adverse events with antibiotic use in hospitalized patients. JAMA Intern Med 177: 1308-1315. [crossref]
7. Huo TI (2010) The first case of multidrug-resistant NDM−1-harboring Enterobacteriaceae in Taiwan: here comes the superbacteria! J Chin Med Assoc 73: 557-558. [crossref]
8. Ferri M, Ranucci E, Romagnoli P, Giaccone V (2017) Antimicrobial resistance: A global emerging threat to public health systems. Crit Rev Food Sci 57: 2857-2876. [crossref]
9. Tan L, Li LY, Ashbolt N, Wang XL, Cui YX, et al. (2018) Arctic antibiotic resistance gene contamination, a result of anthropogenic activities and natural origin. Sci Total Environ 621: 1176-1184. [crossref]
10. Tong S, Pan J, Lu S, Tang J (2018) Patient compliance with antimicrobial drugs: a Chinese survey. Am J Infect Control 46: E25-E29. [crossref]
11. Alygizakis NA, Gago-Ferrero P, Borova VL, Pavlidou A, Hatzianestis I, et al. (2016) Occurrence and spatial distribution of 158 pharmaceuticals, drugs of abuse and related metabolites in offshore seawater. Sci Total Environ 541: 1097-1105. [crossref]
12. Casanova LM, Sobsey MD (2016) Antibiotic-resistant enteric bacteria in environmental waters. Water 8: 561-567.
13. Zhang LH, He YW, Chen M, Gao M, Qiu TL, et al. (2016) Pollution characteristics of antibiotic resistant bacteria from atmospheric environment of animal feeding operations. Huan Jing Ke Xue 37: 4531-4537. [crossref]
14. Jiménez TM, Ferreira IJ, da Silva DS, Guimarães Barrocas PR, Saggioro EM (2018) Removal of contaminants of emerging concern (CECs) and antibiotic resistant bacteria in urban wastewater using UVA/TiO2/H2O2 photocatalysis. Chemosphere 210: 449-457. [crossref]
15. Kerrigan JF, Sandberg KD, Engstrom DR, LaPara TM, Arnold WA (2018) Small and large-scale distribution of four classes of antibiotics in sediment: association with metals and antibiotic resistance genes. Environ Sci Process Impacts 20: 1167-1179. [crossref]
16. McConnell MM, Truelstrup HL, Jamieson RC, Neudorf KD, Yost CK, et al. (2018) Removal of antibiotic resistance genes in two tertiary level municipal wastewater treatment plants. Sci Total Environ643: 292-300. [crossref]
17. Karthikeyan KG, Meyer MT (2006) Occurrence of antibiotics in wastewater treatment facilities in Wisconsin, USA, Sci Total Environ 361: 196-207. [crossref]
18. Dinh QT, Moreau-Guigon E, Labadie P, Alliot F, Teil MJ, et al. (2017) Occurrence of antibiotics in rural catchments. Chemosphere 168: 483-490. [crossref]
19. Dong D, Zhang L, Liu S, Guo Z, Hua X (2016) Antibiotics in water and sediments from Liao River in Jilin Province, China: occurrence, distribution, and risk assessment. Environ Earth Sci 75: 1202.
20. Siedlewicz G, Białk-Bielinska A, Borecka M, Winogradow A, Stepnowski P, et al. (2018).Presence, concentrations and risk assessment of selected antibiotic residues in sediments and near-bottom waters collected from the Polish Coastal Zone in the Southern Baltic Sea-Summary of 3years of studies. Mar Pollut Bull 129: 787-801. [crossref]
21. Barry S (2019) Dangerously high levels of antibiotics found in world’s major rivers, says study. World news global study 1: 1-10.
22. Maycock DS, Watts CD (2011) Pharmaceuticals in drinking water. Encycl Environ Heal 472-484.
23. Fekadu S, Alemayehu E, Dewil R, Van der Bruggen B (2019) Pharmaceuticals in freshwater aquatic environments: a comparison of the african and european challenge. Sci Total Environ 654: 324-337. [crossref]
24. Wang J, Hao J, Liu D, Qin S, Chen C, et al. (2017a) Flower stamen-like porous boron carbon nitride nanoscrolls for water cleaning. Nanoscale 9: 9787-9791. [crossref]
25. Wang X, Wang A, Lu M, Ma J (2017b) Synthesis of magnetically recoverable Fe⁰/graphene-TiO₂ nanowires composite for both reduction and photocatalytic oxidation of metronidazole. Chem Eng J 337: 372–384.
26. Wang W, Zhu Q, Dai Q, Wang X (2017c) Fe doped CeO₂ nanosheets for catalytic oxidation of 1,2-dichloroethane: Effect of preparation method. Chem Eng J 307: 1037-1046.
27. Yang X, Chen Z, Zhao W, Liu C, Qian X, et al. (2021) Recent advances in photodegradation of antibiotic residues in water. Chem Eng J 405: 126806. [crossref]
28. Akyon B, McLaughlin M, Hernández F, Blotevogel J, Bibby K (2019) Characterization and biological removal of organic compounds from hydraulic fracturing produced water. Environ Sci Process 21: 279-290. [crossref]
29. Zhang C, Li Y, Shuai D, Shen Y, Xiong W, et al. (2019) Graphitic carbon nitride (g-C₃N₄)-based photocatalysts for water disinfection and microbial control: A review. Chemosphere 214: 462-479. [crossref]
30. de Souza Santos LV, Meireles AM, Lange LC (2015) Degradation of antibiotics norfloxacin by fenton, UV and UV/H₂O₂. J Environ Manag 154: 8-12. [crossref]
31. Yang L, Zhu YJ, He G, Li H, Tao JC (2022) multifunctional photocatalytic filter paper based on ultralong nanowires of the calcium-alendronate complex for high-performance water purification. ACS Appl Mater Interfaces 14: 9464-9479. [crossref]
32. Zhong Y, Han L, Yin X, Li H, Fang D, et al. (2018) Three dimensional functionalized carbon/tin(IV) sulfide biofoam for photocatalytical purification of chromium(VI)-containing wastewater. ACS Sustain Chem Eng 6: 10660-10667.
33. Alagha O, Ouerfelli N, Kochkar H, Almessiere MA, Slimani Y, et al. (2021) Kinetic modeling for photo-assisted penicillin G degradation of (Mn0.5Zn0.5)[CdxFe2-x]O4 (x ≤ 0.05) nanospinel ferrites. Nanomaterials 11: 970-986. [crossref]
34. The American Society of Health-System Pharmacists. Ofloxacin. Archived from the original on 28 December 2016.
35. Andriole VT (1989) The Quinolones, Academic Press.
36. Johnson JR, Murray AC, Gajewski A, Sullivan M, Snippes P, et al. (2003) Isolation and molecular characterization of nalidixic acid-resistant extraintestinal pathogenic escherichia coli from retail chicken products. Antimicrob Agents Chemother 47: 2161-2168.
37. Kritas S, Ronconi G, Caraffa A, Gallenga C, Ross R, et al. (2020) Mast cells contribute to coronavirus-induced inflammation: new anti-inflammatory strategy. J Biol Regul Homeost Agents 34: 9-14. [crossref]
38. Amini Z, Givianrad MH, Husain SW, Azar PA, Saber-Tehrani M (2020) Cu-S codoping TiO₂/SiO₂ and TiO₂/SiO₂/Fe₃O₄ core-shell nanocomposites as a novel purple LED illumination active photocatalyst for degradation of diclofenac: the effect of different scavenger agents and optimization. Chem Eng Commun. 207: 1536-1553.
39. Sen Gupta PS, Rana MK (2020) Ivermectin, famotidine, and doxycycline: A suggested combinatorial therapeutic for the treatment of COVID-19. ACS Pharmaco Trans Sci3: 1037-1038. [crossref]
40. Tong C, Zhuo X, Guo Y (2011) Occurrence and risk assessment of four typical fluoroquinolone antibiotics in raw and treated sewage and in receiving waters in Hangzhou, China. J Agric Food Chem 59: 7303-7309. [crossref]
41. Xiao Y, Chang H, Jia A, Hu J (2008) Trace analysis of quinolone and fluoroquinolone antibiotics from wastewaters by liquid chromatography–electrospray tandem mass spectrometry. J Chromatogr A 1214: 100-108. [crossref]
42. Anjali R, Shanthakumar S (2019) Insights on the current status of occurrence and removal of antibiotics in wastewater by advanced oxidation processes. J Environ Manage 246: 51-62. [crossref]
43. Liu J, Yin ML, Xiao TF, Zhang CS, Tsang DCW, et al. (2020) Thallium isotopic fractionation in industrial process of pyrite smelting and environmental implications. J Hazard Mater 384: 121378. [crossref]
44. Wang J, Zhuan R (2020) Degradation of antibiotics by advanced oxidation processes: An overview. Sci Total Environ 701: 135023. [crossref]
45. Liu W, Li Z, Kang Q, Wen L (2021a) Efficient photocatalytic degradation of doxycycline by coupling α-Bi₂O₃/g-C₃N₄ composite and H₂O₂ under visible light. Environ Res 197: 110925. [crossref]
46. Liu YP, Lv YT, Guan JF, Khoso FM, Jiang XY, et al. (2021b) Rational design of three-dimensional graphene/graphene oxide-based architectures for the efficient adsorption of contaminants from aqueous solutions. J Mol Liq 343: 117709.
47. Homem V, Santos L (2011) Degradation and removal methods of antibiotics from aqueous matrices-A review. J Environ Manage 92: 2304-2347. [crossref]
48. Bagheri S, TermehYousefi SA, Do TO (2017) Photocatalytic pathway toward degradation of environmental pharmaceutical pollutants: Structure, kinetics and mechanism approach. Catal Sci Technol 7: 4548-4569.
49. Cuerda-Correa EM, Alexandre-Franco MF, Fernandez-Gonzalez C (2020) Advanced oxidation processes for the removal of antibiotics from water, an overview. Water 12: 102-153.
50. Lima VB, Goulart LA, Rocha RS, Steter JR, Lanza MRV (2020) Degradation of antibiotic ciprofloxacin by different AOP systems using electrochemically generated hydrogen peroxide. Chemosphere 247: 125807.
51. Biancullo F, Moreira NFF, Ribeiro AR, Manaia CM, Faria JL, et al. (2019) Heterogeneous photocatalysis using UVA-LEDs for the removal of antibiotics and antibiotic resistant bacteria from urban wastewater treatment plant effluents. Chem Eng J 367: 304-313.
52. Wei Z, Liu J, Shangguan W (2020) A review on photocatalysis in antibiotic wastewater: Pollutant degradation and hydrogen production. Chinese J Catal 41: 1440-1450.
53. Ahmed MM, Brienza M, Goetz V, Chiron S (2014) Solar photo-Fenton using peroxymonosulfate for organic micropollutants removal from domestic wastewater: Comparison with heterogeneous TiO₂ Chemosphere 117: 256-261. [crossref]
54. Wang Y, Sun H, Ang HM, Tade MO, Wang S (2014) Facile synthesis of hierarchically structured magnetic MnO₂/ZnFe₂O₄ hybrid materials and their performance in heterogeneous activation of peroxymonosulfate. ACS Appl Mater Interfaces 6: 19914-19923. [crossref]
55. Ren Y, Lin L, Ma J, Yang J, Feng J, et al. (2015) Sulfate radicals induced from peroxymonosulfate by magnetic ferrospinel MFe₂O₄ (M ¼ Co, Cu, Mn, and Zn) as heterogeneous catalysts in the water. Appl Catal B: Environ 165: 572-578.
56. Mady AH, Baynosa ML, Tuma D, Shim JJ. (2019) Heterogeneous activation of peroxymonosulfate by a novel magnetic 3D gamma-MnO₂@ZnFe₂O₄/rGO nanohybrid as a robust catalyst for phenol degradation. Applied Catalysis B: Environmental 244: 946-956.
57. Zhu S, Xu Y, Zhu Z, Liu Z, Wang W (2020) Activation of peroxymonosulfate by magnetic Co-Fe/SiO₂ layered catalyst derived from iron sludge for ciprofloxacin degradation. Chem Eng J 384: 123298.
58. Ahammad NA, Zulkifli MA, Ahmad MA, Hameed BH, Mohd Din AT (2021) Desorption of chloramphenicol from ordered mesoporous carbonalginate beads: Effects of operating parameters, and isotherm, kinetics, and regeneration studies. J Environ Chem Eng 9: 105015.
59. Jiang Q, Zhang Y, Jiang S, Wang Y, Li H, et al. (2020).Graphene-like carbon sheetsupported nZVI for efficient atrazine oxidation degradation by persulfate activation. Chem Eng J 403: 126309.
60. Shirani Z, Song H, Bhatnagar A (2020) Efficient removal of diclofenac and cephalexin from aqueous solution using Anthriscus sylvestris-derived activated biochar. Sci Total Environ 745: 140789. [crossref]
61. Ahmed MB, Zhou JL, Ngo HH, Guo W (2015) Adsorptive removal of antibiotics from water and wastewater: Progress and challenges. Sci Total Environ 532: 112-126. [crossref]
62. Zhang X, Lou S, Zeng Y (2020a) Facile fabrication of a novel visible light active g-C₃N₄-CoMoO₄ heterojunction with largely improved photocatalytic performance. Mater Lett 281: 128661.
63. Zhang W, Mohamed AR, Ong WJ (2020b) Z-scheme photocatalytic systems for carbon dioxide reduction: where are we now? Angew Chem Int Ed 59: 22894-22915. [crossref]
64. Avcu T, Üner O, Geçgel Ü (2021) Adsorptive removal of diclofenac sodium from aqueous solution onto sycamore ball activated carbon-isotherms, kinetics, and thermodynamic study. Surf Interfaces 24: 101097.
65. Ji Y, Zhang C, Zhang XJ, Xie PF, Wu C (2022) A high adsorption capacity bamboo biochar for CO₂ capture for low temperature heat utilization. Sep Purif Technol 293: 121131.
66. Shahnaz T, Vishnu Priyan V, Pandian S, Narayanasamy S (2021) Use of nanocellulose extracted from grass for adsorption abatement of ciprofloxacin and diclofenac removal with phyto, and fish toxicity studies. Environ Pollut 268: 115494. [crossref]
67. Wang Z, Song L, Wang Y, Zhang XF, Yao J (2021) Construction of a hybrid graphene oxide/nanofibrillated cellulose aerogel used for the efficient removal of methylene blue and tetracycline. J Phys Chem Solids 150: 109839.
68. Cui G, Guo J, Zhang Y, Zhao Q, Fu S, et al. (2019) Chitosan oligosaccharide derivatives as green corrosion inhibitors for P110 steel in a carbon-dioxide-saturated chloride solution. Carbohydr Polym 203: 386-395. [crossref]
69. Phasuphan W, Praphairaksit N, Imyim A (2019) Removal of ibuprofen, diclofenac, and naproxen from water using chitosan-modified waste tire crumb rubber. J Mol Liq 294: 111554.
70. ALOthman ZA, Badjah AY, Alharbi OML, Ali I (2020) Synthesis of chitosan composite iron nanoparticles for removal of diclofenac sodium drug residue in water. Int J Biol Macromol 159: 870-876. [crossref]
71. Yu F, Sun S, Han S, Zheng J, Ma J (2016) Adsorption removal of ciprofloxacin by multi-walled carbon nanotubes with different oxygen contents from aqueous solutions. Chem Eng J 285: 588-595.
72. Zhao H, Liu X, Cao Z, Zhan Y, Shi X, et al. (2016) Adsorption behavior and mechanism of chloramphenicols, sulfonamides, and nonantibiotic pharmaceuticals on multi-walled carbon nanotubes. J Hazard Mater 310: 235-245. [crossref]
73. Bellamkonda S, Thangavel N, Hafeez HY, Neppolian B, Ranga Rao G (2019) Highly active and stable multi-walled carbon nanotubes-graphene-TiO₂ nanohybrid: An efficient non-noble metal photocatalyst for water splitting. Catal Today 321-322: 120-127.
74. Rigueto CVT, Rosseto M, Nazari MT, Ostwald BEP, Alessandretti I, et al. (2021) Adsorption of diclofenac sodium by composite beads prepared from tannery wastesderived gelatin and carbon nanotubes. J Environ Chem Eng 9: 105030.
75. Ahmed MB, Zhou JL, Ngo HH, Guo W, Johir MAH, et al. (2017) Nano-Fe⁰ immobilized onto functionalized biochar gaining excellent stability during sorption and reduction of chloramphenicol via transforming to reusable magnetic composite. Chem Eng J 322: 571-581.
76. Zhao J, Yang X, Liang G, Wang Z, Li S, et al. (2020) Effective removal of two fluoroquinolone antibiotics by PEG-4000 stabilized nanoscale zerovalent iron supported onto zeolite (PZ-NZVI) Sci Total Environ 710: 136289. [crossref]
77. Nguyen LT, Nguyen HT, Pham TD, Tran TD, Chu HT, et al. (2020) UV–visible light driven photocatalytic degradation of ciprofloxacin by N, S co-doped TiO₂: the effect of operational parameters. Top Catal 63: 985-995.
78. Nguyen CH, Tran ML, Van Tran TT, Juang RS (2021) Efficient removal of antibiotic oxytetracycline from water by Fenton-like reactions using reduced graphene oxide-supported bimetallic Pd/nZVI nanocomposites. J Taiwan Inst Chem Eng 119: 80-89.
79. Falyouna O, Idham MF, Maamoun I, Bensaida K, Ashik UPM, et al. (2022a) Promotion of ciprofloxacin adsorption from contaminated solutions by oxalate modified nanoscale zerovalent iron particles. J Mol Liq 359: 119323.
80. Falyouna O, Maamoun I, Bensaida K, Tahara A, Sugihara Y, et al. (2022b) Encapsulation of iron nanoparticles with magnesium hydroxide shell for remarkable removal of ciprofloxacin from contaminated water. J Colloid Interface Sci 605: 813-827. [crossref]
81. Mokhati A, Benturki O, Bernardo M, Kecira Z, Matos I, et al. (2021) Nanoporous carbons prepared from argan nutshells as potential removal agents of diclofenac and paroxetine. J Mol Liq 326: 115368.
82. Shan D, Deng S, Li J, Wang H, He Cet al. (2017) Preparation of porous graphene oxide by chemically intercalating a rigid molecule for enhanced removal of typical pharmaceuticals. Carbon N Y 119: 101-109.
83. Khalil AME, Memon FA, Tabish TA, Salmon D, Zhang S, et al. (2020) Nanostructured porous graphene for efficient removal of emerging contaminants (pharmaceuticals) from water. Chem Eng J 398: 125440.
84. Görmez F, Görmez Ö, Gözmen B, Kalderis D (2019) Degradation of chloramphenicol and metronidazole by electro-fenton process using graphene oxide Fe₃O₄ as heterogeneous catalyst. J Environ Chem Eng 7: 102990.
85. Qiao D, Li Z, Duan J, He X (2020) Adsorption and photocatalytic degradation mechanism of magnetic graphene oxide/ZnO nanocomposites for tetracycline contaminants. Chem Eng J 400: 125952.
86. Yadav S, Asthana A, Singh AK, Chakraborty R, Vidya SS, et al. (2021) Methionine-functionalized graphene oxide/sodium alginate bio-polymer nanocomposite hydrogel beads: synthesis, ısotherm and kinetic studies for an adsorptive removal of fluoroquinolone antibiotics. Nanomaterial,11: 568-593. [crossref]
87. Liu WX, Song S, Ye ML, Zhu Y, Zhao YG, et al. (2022) Nanomaterials with excellent adsorption characteristics for sample pretreatment: a review. Nanomaterials 12: 1845. [crossref]
88. Bao J, Zhu Y, Yuan S, Wang F, Tang H, et al.(2018) Adsorption of tetracycline with reduced graphene oxide decorated with MnFe₂O₄ Nanoscale Res Lett 13: 396-403. [crossref]
89. Tang H, Li W, Jiang H, Lin R, Wang Z, et al. (2019) ZIF-8-derived hollow carbon for efficient adsorption of antibiotics. Nanomaterials 9: 117. [crossref]
90. Wang ZT, Xu JL, Zhou H, Zhang X (2019a) Facile synthesis of Zn(II)-doped g-C₃N₄ and their enhanced photocatalytic activity under visible light irradiation. Rare Met 38: 459-467. [crossref]
91. Wang B, Zhu B, Yun S, Zhang W, Xia C, et al. (2019b) Fast ionic conduction in semiconductor CeO₂-δ electrolyte fuel cells. NPG Asia Mater 11: 51.
92. Masoudi F, Kamranifar M, Safari F, Naghizadeh A (2019) Mechanism, kinetics and thermodynamic of Penicillin G antibiotic removal by silica nanoparticles from simulated hospital wastewater. Desalination Water Treat 169: 333-341.
93. Zeidman AB, Rodriguez-Narvaez OM, Moon J, Bandala ER, et al. (2020) Removal of antibiotics in aqueous phase using silica-based ımmobilized nanomaterials: a review. Environ Technol Innov 20: 101030.
94. Sturini M, Puscalau C, Guerra G, Maraschi F, Bruni G, et al. (2021) Combined layer-by-layer/hydrothermal synthesis of Fe₃O₄@MIL-100(Fe) for ofloxacin adsorption from environmental waters. Nanomaterial 11: 3275. [crossref]
95. Li Y, Gutiérrez Moreno JJ, Song Z, Liu D, et al. (2022) Controlled synthesis of perforated oxide nanosheets with high density nanopores showing superior water purification performance. ACS Appl Mater Interfaces 14: 18513-18524. [crossref]
96. Dehghan A, Mohammadi AA, Yousefi M, Najafpoor AA, Shams M etal (2019) Enhanced kinetic removal of ciprofloxacin onto metal-organic frameworks by sonication, process optimization and metal leaching study. Nanomaterials 9: 1422-1438. [crossref]
97. Sun T, Fan R, Zhang J, Qin M, Chen W, et al. (2021a).Stimuli-Responsive metal–organic framework on a metal–organic framework heterostructure for efficient antibiotic detection and anticounterfeiting. ACS Appl Mater Interfaces 13: 35689-35699. [crossref]
98. Sun L, Cao G, Xu M, Cheng G, Xia D, et al. (2021b) Visible-light-induced peroxymonosulfate activation over ZnFe₂O₄ fine nanoparticles for ofloxacin degradation. Elementa: Science of the Anthropocene 9: 1-10.
99. Luo B, Xu D, Li D, Wu G, Wu Met al. Fabrication of a Ag/Bi3TaO7 plasmonic photocatalyst with enhanced photocatalytic activity for degradation of tetracycline. ACS Appl Mater Interfaces 7: 17061-17069. [crossref]
100. Lu Y, Chu Y, Zheng W, Huo M, Huo H, et al. (2019) Significant tetracycline hydrochloride degradation and electricity generation in a visible-light-driven dual photoelectrode photocatalytic fuel cell using BiVO₄/TiO₂ NT photoanode and Cu₂O/TiO₂ NT photocathode. Electrochim Acta 320: 134617. [crossref]
101. Lu ZY, Ma YL, Zhang JT, Fan NS, Huang BC (2020) A critical review of antibiotic removal strategies: Performance and mechanisms. J Water Process Eng 38: 101681.
102. Yu H, Wang D, Zhao B, Lu Y, Wang X, et al. (2020) Enhanced photocatalytic degradation of tetracycline under visible light by using a ternary photocatalyst of Ag₃PO₄/AgBr/g-C₃N₄ with dual Z-scheme heterojunction. Separ Purif Tech 237: 116365.
103. Chen Y, Yang J, Zeng L, Zhu M (2022) Recent progress on the removal of antibiotic pollutants using photocatalytic oxidation process. Crit Rev Environ Sci Technol 52: 1401-1448.
104. Fu J, Yu J, Jiang C, Cheng B (2018) g-C₃N₄-based heterostructured photocatalysts. Adv Energy Mater 8: 1701503.
105. Ren Y, Zeng D, Ong WJ (2019) Interfacial engineering of graphitic carbon nitride (g-C₃N₄)-based metal sulfide heterojunction photocatalysts for energy conversion: a review Chin J Catal 40: 289-319.
106. Hao Q, Chen T, Wang R, Feng J, Chen D, et al. (2018a) A separation-free polyacrylamide/bentonite/graphitic carbon nitride hydrogel with excellent performance in water treatment. J Clean Prod 197: 1222-1230.
107. Hao X, Zhou J, Cui Z, Wang Y, Zou, Z (2018b) Zn-vacancy mediated electron-hole separation in ZnS/g-C₃N₄ heterojunction for efficient visible-light photocatalytic hydrogen production. Appl Catal B 229: 41-51.
108. Mestre AS, Carvalho AP (2019) Photocatalytic degradation of pharmaceuticals carbamazepine, diclofenac, and sulfamethoxazole by semiconductor and carbon materials: a review. Molecules 24: 3702. [crossref]
109. Low J, Jiang C, Chen B, Wageh S, Al-Ghamdi AA, et al. (2017) A review of direct Z-scheme photocatalysts. Small Methods 1: 1700080.
110. Chen Y, Fan ZX, Zhang ZC, Niu WX, Li CL, et al. (2018) Two-dimensional metal nanomaterials: synthesis, properties, and applications. Chem Rev 118: 6409-6455.
111. Xu Q, Zhang L, Yu J, Wageh S, Al-Ghamdi AA, et al. (2018) Direct Z-scheme photocatalysts: principles, synthesis, and applications. Mater Today 21: 1042-1063.
112. Huang D, Chen S, Zeng G, Gong X, Zhou CY, et al. (2019) Artificial Z-scheme photocatalytic system: What have been done and where to go? Coord Chem Rev 385: 44-80.
113. Yu W, Xu D, Peng T (2015) Enhanced photocatalytic activity of g-C₃N₄ for selective CO₂ reduction to CH₃OH via facile coupling of ZnO: a direct Z-scheme mechanism. J Mater Chem A 3: 19936-19947.
114. Liu X, Pang F, He M, Ge J (2017a) Confined reaction inside nanotubes: New approach to mesoporous g-C₃N₄ Nano Res 10: 3638-3647.
115. Liu D, Zhang M, Xie W, Sun L, Chen Y, et al. (2017b) Porous BN/TiO2 hybrid nanosheets as highly efficient visible-light-driven photocatalysts. Appl Catal B-Environ 207: 72-78.
116. Liu S, Chen F, Li S, Peng X, Xiong Y (2017c) Enhanced photocatalytic conversion of greenhouse gas CO₂ into solar fuels over g-C₃N₄ nanotubes with decorated transparent ZIF-8 nanoclusters. Appl Catal B 211: 1-10.
117. Cui L, Ding X, Wang Y, Shi H, Huang L, et al. (2017) Facile preparation of Z-scheme WO₃/g-C₃N₄ composite photocatalyst with enhanced photocatalytic performance under visible light. Appl Surf Sci B 391: 202-210.
118. Gebreslassie G, Bharali P, Chandra U, Sergawie A, Boruah PK, et al. (2019a/) Novel g-C₃N₄/graphene/NiFe₂O₄ nanocomposites as magnetically separable visible light driven photocatalysts. J Photochem Photobiol A: Chem 382: 111960.
119. Gebreslassie G, Bharali P, Chandra U, Sergawie A, Boruah PK, et al. (2019b) Hydrothermal synthesis of g-C₃N₄/NiFe₂O₄ nanocomposite and its enhanced photocatalytic activity. Appl Organomet Chem 33: e5002.
120. Wu J, Hu J, Qian H, Li J, Yang R, et al. NiCo/ZnO/g-C₃N₄ Z-scheme heterojunction nanoparticles with enhanced photocatalytic degradation oxytetracycline. Diamond Relat Mater 121: 108738.
121. Qu Z, Jing Z, Chen X, Wang Z, Ren H, et al. (2023) Preparation and photocatalytic performance study of dual Z-scheme Bi₂Zr₂O₇/g-C₃N₄/Ag₃PO₄ for removal of antibiotics by visible-light. J Environ Sci 125: 349-361. [crossref]
122. Zhao H, Yu H, Quan X, Chen S, Zhang Y, et al. (2014) Fabrication of atomic single layer graphiticC₃N₄ and its high performance of photocatalytic disinfection under visible light irradiation. Appl Catal B 152-153: 46-50.
123. Mamba G, Mishra AK (2016) Graphitic carbon nitride (gC₃N₄) nanocomposites: A new and exciting generation of visible light driven photocatalysts for environmental pollution remediation. Appl Catal B 198: 347-377.
124. Darkwah WK, Ao Y (2018) Mini review on the structure and properties (photocatalysis), and preparation techniques of graphitic carbon nitride nano-based particle, and its applications. Nanoscale Res Lett 388: 1-15.
125. Oh WD, Chang VWC, Hu ZT, Goei R, Lim TT.(2017) Enhancing the catalytic activity of g-C₃N₄ through Me doping (Me = Cu, Co and Fe) for selective sulfathiazole degradation via redox-based advanced oxidation process. Chem Eng J 323: 260-269.
126. Dai Y, Gu Y, Bu Y (2020) Modulation of the photocatalytic performance of g-C3N4 by two-sites co-doping using variable valence metal. Appl Surf Sci 500: 144036.
127. Ye L, Liu J, Jiang Z, Peng T, Zan L (2013) Facets coupling of BiOBr-g-C₃N₄ composite photocatalyst for enhanced visible-light-driven photocatalytic activity. Appl Catal B 142-143: 1-7.
128. Tian N, Huang H, He Y, Guo Y, Zhang T, et al. (2015a) Mediator-free direct Z-scheme photocatalytic system: BiVO₄/g-C₃N₄ organic–inorganic hybrid photocatalyst with highly efficient visible-light-induced photocatalytic activity. Dalton Trans 44: 4297-4307.
129. Tian N, Huang H, Liu C, Dong F, Zhang T, et al. (2015b) In situ co-pyrolysis fabrication of CeO₂/g-C₃N₄ n-n type heterojunction for synchronously promoting photo-induced oxidation and reduction properties. J Mater Chem A 3: 17120-17129.
130. Nithya R, Ayyappan S.(2020) Novel exfoliated graphiticC₃N₄ hybridised ZnBi₂O₄ (g-C₃N₄/ZnBi₂O₄) nanorods for catalytic reduction of 4-Nitrophenol and its antibacterial activity J Photochem Photobiol Chem 398: 112591.
131. Tian N, Huang H, Wang S, Zhang T, Du X, et al. (2020) Facet-charge-induced coupling dependent interfacial photocharge separation: A case of BiOI/g-C₃N₄ p-n junction. Appl Catal B 267: 118697
132. Montini T, Melchionna M, Monai M, Fornasiero P (2016) Fundamentals and catalytic applications of CeO₂-based materials. Chem Rev 116: 5987-6041. [crossref]
133. Fauzi AA, Jalila AA, Hassan NS, Aziz FFA, Azami MS, et al. (2022) A critical review on relationship of CeO₂-based photocatalyst towards mechanistic degradation of organic pollutant. Chemosphere 286: 131651. [crossref]
134. Kashmery HA, El-Hout SI, Zak ZI (2022) Fast photocatalytic oxidation of ciprofloxacin over Co₃O₄@CeO₂ heterojunctions under visible-light. J Taiwan Inst Chem Eng 140: 104563.
135. Rajeshkumar S, Naik P (2018) Synthesis and biomedical applications of Cerium oxide nanoparticles-A Review. Biotechnol Rep 17: 1-5.
136. Latif MM, Haq AU, Amin F, Ajaz-un-Nabi M, Khan I, et al. (2021) Synthesis and antimicrobial activities of manganese (Mn) and iron (Fe) co-doped cerium dioxide (CeO₂) nanoparticles. Physica B Condens Matter 600: 412562.
137. Parvathy S, Subramanian P, Karthick SA, Subbaiya R (2022) The structural, optical, antimicrobial and anticancer properties of biocompatible astaxanthin coated ZnO and CeO₂ Mater Lett 312: 131669.
138. Sharma D, Sastsangi VR, Shrivastav R, Waghmare UV, Dass S. (2016).Understanding the photoelectrochemical properties of nanostructured CeO₂/CuO₂ heterojunction photoanode for efficient photoelectrochemical water splitting. Int J Hydrog Energy 41: 18339-18350.
139. Sharma D, Mehta BR (2018) Nanostructured TiO₂ thin films sensitized by CeO₂ as an inexpensive photoanode for enhanced photoactivity of water oxidation. J Alloys Compd 749: 329-335.
140. Yin D, Zhao F, Zhang L, Zhang X, Liu Y, et al. (2016) Greatly enhanced photocatalytic activity of semiconductor CeO₂ by integrating with upconversion nanocrystals and graphene. RSC Adv 6: 103795-103802.
141. Li R, Li C, Yin S, Fu C, Satom T (2012) Synthesis of C-N Co-doped nano-CeO₂ and dye degradation under compact fluorescent lamp irradiation. J Nanosci Nanotechnol 12: 2797-2801. [crossref]
142. Channei D, Inceesungvorn B, Wetchakun N, Ukritnukun S, Nattestad A, et al. Photocatalytic degradation of methyl orange by CeO₂ and Fe-doped CeO₂ films under visible light irradiation. Sci Rep 4: 5757. [crossref]
143. Huang K, Li YH, Liang C, Xu X, Zhou YF, et al. (2014) One-step synthesis of reduced graphene oxide-CeO₂ nanocubes composites with enhanced catalytic activity. Mater Lett 124: 223-226.
144. Qu X, Xie D, Gao L, Du F (2014) Synthesis and photocatalytic activity of TiO₂ /CeO₂ core-shell nanotubes. Mater Sci Semicond Process 26: 657-662.
145. Saranya J, Ranjith KS, Saravanan P, Mangalaraj D, Kumar RTR (2014) Cobalt-doped cerium oxide nanoparticles: Enhanced photocatalytic activity under UV and visible light irradiation. Mater Sci Semicond Process 26: 218-224.
146. Zhang C, Zhang X, Wang Y, Xie S, Liu Y, et al. (2014) Facile electrochemical synthesis of CeO₂ hierarchical nanorods and nanowires with excellent photocatalytic activities. New J Chem 38: 2581-2586.
147. Issarapanacheewin S, Wetchakun K, Phanichphant S, Kangwansupamokon W, Wetchakun N (2015) A novel CeO₂ /Bi₂WO₆ composite with highly enhanced photocatalytic activity. Mater Lett 156: 28-31.
148. Mittal M, Gupta A, Pandey OP (2018) Role of oxygen vacancies in Ag/Au doped CeO₂ nanoparticles for fast photocatalysis. Sol Energy 165: 206-216.
149. Xiu Z, Xing Z, Li Z, Wu X, Yan X, et al. (2018) Ti³⁺-TiO₂ /Ce³⁺-CeO₂ nanosheet heterojunctions as efficient visible-light-driven photocatalysts. Mater Res Bull 100: 191-197.
150. Toro RG, Malandrino G, Fragala IL (2004) Relationship between the nanostructures and the optical properties of CeO₂ thin films. J Phys Chem B 108: 16357-16364.
151. Chen HI, Chang HY (2005) Synthesis and characterization of nanocrystalline cerium oxide powders by two-stage non-isothermal precipitation. Solid State Commun 133: 593-598.
152. Avellaneda CO, Berton MAC, Bulhões LOS (2008) Optical and electrochemical properties of CeO₂ thin film prepared by an alkoxide route. Sol Energy Mater Sol Cells 92: 240-244.
153. Wongkaew A (2008) Effect of cerium oxide and zirconium oxide to activity of catalysts. Chiang Mai J Sci 35: 156-162.
154. Zhang N, Zhang G, Chong S, Zhao H, Huang T, et al. (2018) Ultrasonic impregnation of MnO₂/CeO₂ and its application in catalytic sono-degradation of methyl orange J Environ Manag 205: 134-141. [crossref]
155. Aboutaleb WA, El-Salamony RA (2019) Effect of Fe₂O₃ -CeO₂ nanocomposite synthesis method on the Congo red dye photodegradation under visible light irradiation. Mater Chem Phys 236: 121724.
156. Munoz-Batista MJ, Gomez-Cerezo MN, Kubacka A, Tudela D, Fernandez-Garcia M (2014) Role of interface contact in CeO₂-TiO₂ photocatalytic composite materials. ACS Catal 4: 63-72.
157. Kusmierek E (2020) A CeO₂ semiconductor as a photocatalytic and photoelectrocatalytic material for the remediation of pollutants in ındustrial wastewater: a review. Catalysts 10: 1435-1489.
158. Baird RB, Eaton AD, Rice EW (2017) Standard Methods for the Examination of Water and Wastewater.
159. Olthof M, Eckenfelder WW (1976) Coagulation of textile wastewater. Text Chem Color8 : 18-22.
160. Eckenfelder WW (1989) Industrial Water Pollution Control (2nd ed), Signapore: McGraw-Hill Inc.
161. Lange B (1994) LUMISmini, Operating Manual. Dusseldorf, Germany.
162. Lange B (2010) Vibrio fischeri -Microtox LCK 491 kit. Germany.
163. Lange B (1996) LUMIXmini type luminometer. Dusseldorf.
164. Zar JH (1984) Biostatistical analysis, Prentice-Hall, Englewood Cliffs.
165. Statgraphics Centurion XV, sofware, StatPoint Inc, Statgraphics Centurion XV, Herndon, VA, USA, 2005.
166. Eddy DR, Puri FN, Noviyanti AR (2015) Synthesis and photocatalytic activity of silica-based sand quartz as the supporting TiO₂ Proced Chem 17: 55-58.
167. Li H, Hu J, Meng Y, Su J, Wang X (2017a) An investigation into the rapid removal of tetracycline using multilayered graphene-phase biochar derived from waste chicken feather. Sci Total Environ 603-604: 39-48. [crossref]
168. Li X, Zhu W, Lu X, Zuo S, Yao C, et al. (2017b) Integrated nanostructures of CeO₂ /attapulgite/g-C₃N₄ as efficient catalyst for photocatalytic desulfurization: Mechanism, kinetics and influencing factors. Chem Eng J 326: 87-98. [crossref]
169. Vignesh S, Suganthi S, Sundar JK, Raj V (2019) Construction of α-Fe₂O₃/CeO₂ decorated g-C₃N₄ nanosheets for magnetically separable efficient photocatalytic performance under visible light exposure and bacterial disinfection. Appl Surf Sci 488: 763-777.
170. Huang L, Li Y, Xu H, Xu Y, Xia J, et al. (2013) Synthesis and characterization of CeO₂/g-C₃N₄ composites with enhanced visible-light photocatalytic activity. RSC Adv 3: 22269-22279.
171. Yuan Y, Huang GF, Hu WY, Xiong DN, Zhou BX, et al. (2017) Construction of g-C₃N₄ /CeO₂/ZnO ternary photocatalysts with enhanced photocatalytic performance. J Phys Chem Solids 106: 1-9.
172. Liu W, Zhou J, Hu Z (2019) Nano-sized g-C3N4 thin layer @ CeO2 sphere core-shell photocatalyst combined with H₂O₂ to degrade doxycycline in water under visible light irradiation. Sep Purif Technol 227: 115665.
173. Murali A, Sarswat PK, Free ML (2010) Minimizing electron-hole pair recombination through band-gap engineering in novel ZnO-CeO₂-rGO ternary nanocomposite for photoelectrochemical and photocatalytic applications. Environ Sci Pollut Res 27: 25042-25056.
174. Dixit D, Verma A, Gupta S, Bansal P (2016) Assessment of solar photocatalytic degradation and mineralization of amoxicillin trihydrate (AMT) using slurry and fixed-bed batch reactor: efficacy of parabolic trough collector. RSC Adv 6: 36109-361017.
175. Ferri M, Ranucci E, Romagnoli P, Giaccone V (2017) Antimicrobial resistance: A global emerging threat to public health systems. Crit Rev Food Sci 57: 2857-2876. [crossref]
176. Guo X, He S, Meng Z, Wang Y, Peng Y (2022) Ag@ZIF-8/g-C₃N₄ Z-scheme photocatalyst for the enhanced removal of multiple classes of antibiotics by integrated adsorption and photocatalytic degradation under visible light irradiation. The Royal Society of Chemistry, RSC Adv 12: 17919-17931.
177. Lei W, Zhang T, Gu L, Liu P, Rodriquez JA, et al. (2015) Surface-structure sensitivity of CeO₂ nanocrystals in photocatalysis and enhancing the reactivity with nanogold. ACS Catal 5: 4385-4393.
178. Pellerito A, Ameen SM, Micali M, Caruso G (2018) Antimicrobial substances for food packaging products: the current situation. J AOAC Int 101(4): 942-947.
179. Sizykh M, Batoeva M, Matafonova G (2023) Enhanced catalyst-free degradation and mineralization of ceftriaxone by UV/H₂O₂ and UV/S₂O₈²⁻processes using KrCl excilamp (222 nm) J Photochem Photobiol A Chem 436: 114357.
180. Ye Z, Li J, Zhou M, Wang H, Ma Y, et al. (2016) Well-dispersed nebula-like ZnO/CeO₂@HNTs heterostructure for efficient photocatalytic degradation of tetracycline. Chem Eng J 304: 917-933.
181. Zhao Y, Liang X, Wang Y, Shi H, Liu E, et al. (2018) Degradation and removal of Ceftriaxone sodium in aquatic environment with Bi₂WO₆/g-C₃N₄ J Colloid Interface Sci 523: 7-17. [crossref]
182. Zhou JB, Liu W, Cai WQ (2019) The synergistic effect of Ag/AgCl@ZIF-8 modified g-C₃N₄ composite and peroxymonosulfate for the enhanced visible-light photocatalytic degradation of levofloxacin. Sci Total Environ 696: 133962. [crossref]