**Contents**


## **About the Editor**

**Annalisa Vacca**, a graduate in Chemical Engineering and PhD in Industrial Engineering, was granted the position of full professor in the field of "Fundamentals of chemical technology" in 2022. She carries out didactic and scientific research in the Department of Mechanical, Chemical and Materials Engineering at the University of Cagliari. Her research activities are focused on the field of electrochemical engineering applied to the study of processes for environmental remediation and energy conversion. In particular, her studies cover key aspects such as the catalytic activity of electrode materials and the identification of reaction mechanisms, as well as practical aspects such as the design and characterization of electrochemical reactors.

### *Editorial* **Materials and Processes for Photocatalytic and (Photo)Electrocatalytic Removal of Bio-Refractory Pollutants and Emerging Contaminants from Waters**

**Annalisa Vacca**

Dipartimento di Ingegneria Meccanica, Chimica e dei Materiali, Università di Cagliari, Piazza D'armi, 09123 Cagliari, Italy; annalisa.vacca@dimcm.unica.it

This volume is focused on materials and processes for the electro- and photoelectrochemical removal of biorefractory pollutants and emerging contaminants from waters to show the importance of electrochemistry and photoelectrochemistry in offering solutions to current environmental problems. In addition, we highlight their interdisciplinarity and emphasize the fundamental and applied aspects of these methods.

The research for innovative methods for removing pollutants from water has grown along with the detection of new contaminants in water bodies, the so-called emerging pollutants (EP), that can affect both flora and fauna and human health [1]: they include products used daily in households, industry, pharmaceuticals and personal care products, gasoline additives, plasticizers and microplastics [2]. Two main issues of EP are their dynamic character, which is also connected to the improvement of detection techniques, and the difficulty of removal by conventional wastewater treatment technologies. Moreover, emerging pollutants constitute a threat—even at a trace level—because their real impact on human health is unknown.

Although there are no discharge limits for most EP up to now, the European Commission has drawn up and implemented a watch list containing several chemical contaminants that must be monitored with the aim to generate high-quality data on their concentrations in the aquatic environment and to support the risk assessments that underpin the identification of priority substances [3].

During recent years, electro- and photoelectrochemical processes have demonstrated their capacity to efficiently oxidize many of these compounds. Starting from the early 1980s, research on the electrochemical methods for treated wastewater has grown significantly, and thousands of papers now appear in the literature. Although several tests demonstrate the effectiveness of pollutant removal from synthetic and real matrices, this technology is still far from full-scale applications. Its TRL (technology readiness level) is between 4 (technology validated in the lab) and 5 (technology validated in a relevant environment) [4].

More recently, photoelectrochemical processes in which electrochemical and photochemical processes are combined has attracted increasing interest, thanks to the synergy of the two processes: the application of a bias potential improves the photochemical process and the electrochemical process is more efficient since the photo-potential generated on the semiconductor allows for the depolarizing of the cell. This is why, in the last two decades, the number of articles on photochemical wastewater treatment has quickly increased, and the publication of these articles in specific journals indicates that the technology is moving from the fundamentals to real applications [5]. Nevertheless, the TRL of the photoelectrochemical treatment of wastewater is still at the lab scale, and much more efforts are required to push this technology toward applications in the field.

Thus, this special issue contributes to this context, addressing the synthesis, characterization, and application of new materials, as well as the study of catalytic processes and reaction kinetics.

**Citation:** Vacca, A. Materials and Processes for Photocatalytic and (Photo)Electrocatalytic Removal of Bio-Refractory Pollutants and Emerging Contaminants from Waters. *Catalysts* **2021**, *11*, 666. https:// doi.org/10.3390/catal11060666

Received: 28 April 2021 Accepted: 21 May 2021 Published: 24 May 2021

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2021 by the author. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

I thank all of the authors for their valuable contribution to this Special Issue and the editorial team at *Catalyst* for their kindness and constant support.

**Funding:** This research received no external funding.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


### *Article* **Enhanced Photocatalytic Activity of Au**/**TiO2 Nanoparticles against Ciprofloxacin**

**Pedro Martins 1,2,\*, Sandro Kappert 3, Hoai Nga Le 3,4, Victor Sebastian 5,6, Klaus Kühn 3, Madalena Alves 1, Luciana Pereira 1, Gianaurelio Cuniberti 3,7,8, Manuel Melle-Franco <sup>9</sup> and Senentxu Lanceros-Méndez 1,10,11,\***


Received: 14 January 2020; Accepted: 11 February 2020; Published: 15 February 2020

**Abstract:** In the last decades, photocatalysis has arisen as a solution to degrade emerging pollutants such as antibiotics. However, the reduced photoactivation of TiO2 under visible radiation constitutes a major drawback because 95% of sunlight radiation is not being used in this process. Thus, it is critical to modify TiO2 nanoparticles to improve the ability to absorb visible radiation from sunlight. This work reports on the synthesis of TiO2 nanoparticles decorated with gold (Au) nanoparticles by deposition-precipitation method for enhanced photocatalytic activity. The produced nanocomposites absorb 40% to 55% more radiation in the visible range than pristine TiO2, the best results being obtained for the synthesis performed at 25 ◦C and with Au loading of 0.05 to 0.1 wt. %. Experimental tests yielded a higher photocatalytic degradation of 91% and 49% of ciprofloxacin (5 mg/L) under UV and visible radiation, correspondingly. Computational modeling supports the experimental results, showing the ability of Au to bind TiO2 anatase surfaces, the relevant role of Au transferring electrons, and the high affinity of ciprofloxacin to both Au and TiO2 surfaces. Hence, the present work represents a reliable approach to produce efficient photocatalytic materials and an overall contribution in the development of high-performance Au/TiO2 photocatalytic nanostructures through the optimization of the synthesis parameters, photocatalytic conditions, and computational modeling.

**Keywords:** Au-TiO2; antibiotics; emergent contaminants; nanocatalyst; photocatalysis; GFN-xTB

#### **1. Introduction**

The resilience of specific emerging pollutants such as pharmaceuticals to the traditional wastewater treatments makes them spread in variable concentrations in surface and groundwater [1]. Dissemination of antibiotics in nature is one of the most significant environmental concerns as they affect biological metabolism and induce the presence of bacterial resistance among drinking water sources [2]. Photocatalysis has received considerable attention from the scientific community as a possible solution to degrade these compounds [3,4].

Typically, the photocatalytic process takes place when a catalyst is UV irradiated and electron-hole pairs are created that will react with H2O, OH−, and O2 to generate oxidizing species such as the hydroxyl radical (OH•), superoxide radical anions (O2•−), and hydrogen peroxide (H2O2). These species will initiate a series of reactions that will degrade pollutants into harmless compounds (e.g., CO2 and H2O).

Photocatalysis presents several advantages when compared with other methods, such as the low cost, and the eco-friendly and straightforward processing conditions [5,6]. Many photocatalysts have been reported in the last decades [7,8]. Among them, titanium dioxide (TiO2) is the most studied and applied in photocatalysis, mainly because of its remarkable optical and oxidizing properties, superhydrophilicity, chemical stability, and durability [9,10]. Despite the compelling advantages of TiO2, there are also some drawbacks. One of the main hurdles is the low spectral activation of TiO2, caused by its wide bandgap (3.0–3.2 eV) excitation that only occurs under radiation in the UV or near the UV region (410–387 nm) [11].

For this reason, solar radiation cannot be efficiently used because only less than 5% of this radiation corresponds to UV [3]. Additionally, the process becomes less cost effective as the UV lamps are required to provide the radiation. Another limitation is the electron-hole pair recombination that decreases the photocatalytic efficiency [12,13].

The research developed in the last decades has been mainly devoted to surpassing those limitations by producing new and more efficient photocatalytic materials. Strategies for metallic and nonmetallic doping, co-doping [14,15], dye sensitization, semiconductor combination, co-catalyst loading, and nanocomposite materials [16,17] have been used and tested. These approaches allow us to reduce the electron-hole recombination rate and enhance the absorption of visible radiation of TiO2 by introducing intermediate energy levels inside the bandgap [18]. In this scope, several works have reported the functionalization of TiO2 nanoparticles surfaces with metals such as Au [19], Cu [20], Co [21], and Ag [22]. When irradiated, noble metals nanoparticles at the TiO2 surface can receive electrons and prevent the recombination of the photo-generated electron-hole pairs [23,24].

Metals such as Au and Ag can increase visible light absorption due to the surface plasmon resonance effect [25,26]. Gold (Au) nanoparticles have attracted considerable attention, mainly because they possess exceptional stability, nontoxicity, and biocompatibility [3]. Their properties are highly dependent on the size and shape of the nanoparticles, allowing a broad range of applications [27,28]. For instance, the literature shows that gold nanoparticles in the range of 5 to 10 nm present an enhanced catalytic activity [29,30]. In this sense, some works focused on the photocatalytic activity of Au/TiO2 nanocomposite have been published, including interesting review articles [3,29,31].

Different physical-chemical techniques have been exploited to produce Au/TiO2 nanocomposites with enhanced catalytic properties. For instance, chemical vapor deposition [32], sol-gel [33], spray pyrolysis [34], electrophoretic approach [35], deposition-precipitation (DP) [36], deposition-precipitation using urea [37], impregnation [38], hybridization [39], and surface functionalization [40], among others [41,42]. However, many of these techniques are time-consuming, and few of them have focused on the optimization of the nanocomposite and the computational modeling of its nanostructure. Thus, this work focused on the optimization of a DP, converting the Au/TiO2 nanocomposite production into a cost-effective and straightforward technique, with enhanced photocatalytic activity, under UV and visible radiation. The method optimization aims for cost reduction, using the lowest Au loading that endows visible spectra photocatalytic activity to the nanocomposite. The computational studies

provide further information about the electronic mechanism behind the enhanced photocatalytic activity of the Au/TiO2 nanocomposite, as well as the interaction with the target compound.

The target compound is the fluoroquinolone ciprofloxacin (CIP) (chemical formula in Supplementary Material, Figure S1), belonging to a class of synthetic broad-spectrum antibiotics [43], which is mostly used in medicine (e.g., tuberculosis, pneumonia, or digestive disorders). It is also one of the most prescribed fluoroquinolones in the world and studies has shown its presence in potable water and wastewater, as well as in sewage sludge at variable concentrations from milligrams to nanograms per liter [2,44].

In this work, photocatalytic efficiency during the degradation of CIP under UV and visible illumination was assessed. To the best of our knowledge, this is the first work that combines an optimization process of Au/TiO2 nanocomposite with photocatalytic experiments for CIP degradation and computational modeling that addresses the interaction between Au and TiO2 nanoparticles, as well as the interaction of CIP with the produced nanocomposites.

#### **2. Results and Discussion**

#### *2.1. Nanocomposite Characterization*

The Au/TiO2 nanocomposites were produced by nanoprecipitation method, and the temperature (25, 60, and 80 ◦C) and the Au loading (ranging from 0.025 to 0.5 wt. %) were changed to understand how these parameters affect the morphology of the nanocomposites and relate it to the photocatalytic efficiency. In this sense, scanning transmission electron microscopy-high-angle annular dark-field imaging (STEM-HAADF) analysis was performed, and the micrographs of the different nanocomposites are displayed in Figure 1.

**Figure 1.** Scanning transmission electron microscopy-high-angle annular dark-field imaging micrographs of Au/TiO2 nanocomposites synthesized with different Au loadings at 60 ◦C (**a**–**c**), and Au/TiO2 nanocomposites obtained at different temperatures with an Au loading of 0.05 wt. % (**d**–**f**).

The Au loading study (Figure 1a–c) was assessed producing different nanocomposites using the same experimental conditions (temperature = 60 ◦C) and changing the loading of gold exclusively, from 0.025 to 0.5 wt. %. The STEM-HAADF micrographs show that for the sample with 0.025 wt. % of Au (Figure 1a), the presence of Au nanoparticles over the surface of the TiO2 nanoparticles was almost inexistent (Figure 1a). With the increase of Au loading to 0.05 wt. % (Figure 1b), it was possible to observe a homogeneous distribution of predominantly small Au nanoparticles (bright contrast

nanoparticles below 5 nm in diameter) over the TiO2 nanoparticles. Similar results were obtained for 0.1 wt. % (data not shown). For the concentrations of 0.25 and 0.5 wt. % (Figure 1a–c), agglomerates of Au over theTiO2 nanoparticles (brightest areas of the micrograph) were identified as well as large Au nanoparticles. Analogously, the effect of temperature on the synthesis product was also performed maintaining all the synthesis parameters (Au loading = 0.05 wt. % yielded a homogeneous distribution and size of Au nanoparticles) and changing the temperature of the different samples. STEM-HAADF images (Figure 1d–f) indicate that although the used Au loading was the same in the three temperatures tested when the nanocomposite was synthesized at 80 ◦C, larger Au nanoparticles appeared more frequently on the nanocomposite (Figure 1f). Conversely, at lower temperatures (25 and 60 ◦C), the Au nanoparticles size was smaller (Figure 1d,e).

The study of the effect of Au loading and temperature in the nanocomposites morphology indicates that the samples produced at 60 ◦C and with an Au loading of 0.05 wt. % possessed the more homogeneous distribution and size of Au nanoparticles. In this way, a more detailed STEM-HAADF analysis (Figure 2) was performed on this sample. Figure 2a,b reveal a homogeneous dispersion of Au nanoparticles (white arrows) over TiO2 nanoparticles' surface. The representation of the sphere-like shape of Au nanoparticles in Figure 2c, where an high-resolution scanning transmission electron microscopy – high-angle annular dark field shows that single-crystal nanoparticles with high crystallinity were produced by the proposed method. Size distribution, ranging from 1 to 7 nm, and the average size of 3.2 ± 1.13 nm (Figure 2d), were quantified using Image J software applied to 400 nanoparticles. The size distribution of Au nanoparticles for synthesis at 25 ◦C and 80 ◦C is provided in Supplementary Material (Figure S2). All the images show Au nanoparticles with similar sizes, which is in good agreement with the size distribution histogram that presents a sharp size distribution.

**Figure 2.** STEM-HAADF micrographs of Au/TiO2 nanocomposites (produced at 60 ◦C and Au loading of 0.05 wt. %) at different scales (**a**) 50 and (**b**) 200 nm; detail of Au nanoparticle over TiO2 nanoparticles' surface and single Au nanoparticle amplification (inset) (**c**); size distribution of 400 Au nanoparticles with the respective average size (**d**).

The STEM-HAADF- energy-dispersive X-ray spectroscopy (EDX) measurements allowed us to identify the elements present in the Au/TiO2 sample in two different points, 1 and 2 (signaled in Figure 3a). STEM-HAADF-EDX spectra in Figure 3b in point 1 indicate the presence of Au and Cu (copper), which can be respectively addressed to Au nanoparticles and copper grid. In point 2, the signatures of Ti (titanium) and O (oxygen) were identified, corresponding to TiO2 nanoparticles. Thus, EDX measurements confirmed the presence of all the elements of the Au/TiO2 nanocomposite.

**Figure 3.** The STEM-HAADF- energy-dispersive X-ray spectroscopy (EDX) image of Au/TiO2 nanocomposites with the identification of the measured points: Au (1) and TiO2 (2) (**a**), EDX spectra with elemental identification (Au, Ti, O, and C) for points 1 and 2 (**b**). The Au/TiO2 nanocomposite synthesized at 60 ◦C and with an Au loading of 0.05 wt. % was used.

X-ray diffraction was performed to assess the crystal structure of the pure TiO2 nanoparticles and Au/TiO2 nanocomposite, Figure 4a. Both samples show the typical reflexes from anatase (25.3◦, 37.8◦, and 48.0◦) and rutile (27.49◦). There was no significant difference between the intensities or positions of the reflexes from both samples. Moreover, no reflexes of Au were detected, which can be explained by the low amount of Au present in the nanocomposite (below detection limit). Figure 4b shows the study of hydrodynamic size for TiO2 and Au/TiO2 nanocomposites obtained by dynamic light scattering (DLS). The results indicated nanoparticles diameters of 1023 nm and 342 nm, for the pristine TiO2 and the Au/TiO2 nanocomposites, respectively. The results suggest that the presence of Au nanoparticles over TiO2 nanoparticles surface may prevent the formation of nanoparticles' aggregates. On the other hand, the size distribution was broader for the nanocomposites regarding the pristine TiO2. Previous work equally showed that the presence of erbium (Er) on TiO2 nanoparticles contributed to reducing the hydrodynamic size when compared with bare TiO2 [15].

The zeta potential was studied at different pH values (3, 5, 7, 9, and 11) for TiO2 and Au/TiO2 samples and the results are displayed in Figure 4c. The pristine and the Au/TiO2 presented very similar profiles, with higher zeta potential values ≈ |20| mV for pH below 3 and 9. These data were in good agreement with the literature [45], with positive zeta potential values for acidic conditions and negative values for basic pH. The more significant difference between the two samples occurred at pH = 7, with the nanocomposite presenting higher zeta potential values than the pure TiO2. Higher zeta potential values mean that nanoparticles possess higher periphery surface charge, which promotes nanoparticles' repulsions, avoiding aggregates' formation and enhanced stability [46]. In this context, and relating it with DLS-obtained results, the smaller hydrodynamic size was probably obtained for the Au/TiO2 because repulsions endowed by Au on TiO2 nanoparticles surface prevented the formation of the aggregates.

**Figure 4.** X-ray diffraction reflexes of pristine TiO2 and Au/TiO2 nanocomposite and identification of the representative peaks for anatase (A) and rutile (R) phases (**a**); dynamic light scattering, intensity size distribution of the pristine TiO2 and the Au/TiO2 nanocomposite and respective Z-average hydrodynamic size (**b**); zeta potential measurements, performed at different pHs (3, 5, 7, 9, and 11) for pristine TiO2 nanoparticles and Au/TiO2 nanocomposite (**c**); UV–vis reflectance spectra of pristine TiO2 and Au/TiO2 and (inset) the estimation of the bandgap for both samples at (F(R))1/2 = 0 (**d**). The Au/TiO2 nanocomposite synthesized at 60 ◦C and with an Au loading of 0.05 wt. % was used.

To understand the differences in the photocatalytic performance of TiO2 and Au/TiO2 nanocomposite, the optical properties of these materials were studied by UV-visible diffuse reflectance spectra (DRS), depicted in Figure 4d. In the visible range (400–800 nm), the pure TiO2 nanoparticles reflect the radiation almost entirely (≈ 95%). However, the nanocomposite displays reflectance below 64% for the same range. Additionally, a minimum reflectance (≈ 44%) was obtained at 545 nm, indicating a maximum of absorbance band that can be associated with the surface plasmon of Au nanoparticles, typically in the wavelength range between 520 and 560 nm [47,48]. These results show that the nanocomposite presented a broad absorbance spectrum when compared to the pristine TiO2 nanoparticles, which is also consistent with the purple/pink color exhibited by the produced nanocomposite. In the ultraviolet range (200–400 nm), both samples showed similar behavior.

From DRS spectra it was possible to estimate the band gap, shown in the inset graph of Figure 4d, for pure TiO2 and Au/TiO2 nanocomposite was converting the reflectance to Kubelka–Munk units through Equation (1) and Equation (2). The obtained values show that the nanocomposites possessed a lower bandgap (2.84 eV) than the pristine TiO2 nanoparticles (2.96 eV). The decrease of the bandgap in Au/TiO2 was related to the shift absorption to longer wavelengths. Similar results have been reported in the literature [49,50].

#### *2.2. Nanocomposites' Optimization and Photocatalytic Experiments*

The photocatalytic activity of all the produced Au/TiO2 nanocomposites was assessed by monitoring the degradation of CIP under artificial UV and visible irradiations. Process conditions were varied depending on the studying purposes.

#### Nanocomposite Optimization

As gold is a noble metal, cost-effectiveness should be considered, and the amount of gold used in the nanocomposite is one of the most paramount parameters. In this study, Au loading was varied by using different concentrations of the gold precursor. The tested Au loadings were 0.025, 0.05, 0.1, 0.25, and 0.5 wt. %. These nanocomposites were employed for the photocatalytic degradation of CIP under both UV and simulated visible radiation.

Figure 5a shows the data of photocatalytic experiments under UV light. Accordingly, all produced samples and the pristine TiO2 used as a control showed photocatalytic activity, proven by the decrease of CIP concentration along with the irradiation time. As confirmed by the diffuse reflectance spectroscopy (Section 2.1), the bandgap of the nanocomposites was 2.84 eV, corresponding to the wavelength of 437 nm. Here, the used UV lamp had the mode wavelength of 365 nm, which was shorter than the bandgap. It means that the photon energy was adequate to excite the photocatalytic materials, and photocatalytic reaction occurred in all experiments. Pristine TiO2 was compared with the synthesized photocatalysts. After 30 min, 77% of CIP was degraded in the presence of pristine TiO2, whereas higher degradation of 80–90% was achieved in the same time of irradiation, using the synthesized photocatalysts, 0.05 wt. %. This efficiency can be assigned to the presence of gold particles on the surface of the photocatalysts, confirmed by the TEM and EDX characterization (Section 2.1). The further quantitative inspection was obtained using the Langmuir–Hinshelwood kinetics (Equation (3)), and data are shown in Table 1. The apparent reaction rate constant k of the experiment with the bare TiO2 was found to be 0.047 min−1, while the decoration with gold particles improved the photocatalytic activity by 2–3 times. As predicted, in the presence of gold, the excited electrons may be conducted to the gold particles, and the electron-hole recombination may be reduced, which prolongs the lifetime of generated holes [51,52].

Consequently, the photocatalytic activity of the composites increased. Additionally, the increase of the used chloroauric acid concentration might induce a more significant number of gold particles distributed on the TiO2 surface. In other words, the number of electron absorption centers was increased, which explains the increase of k from 0.078 to 0.131 min−<sup>1</sup> when increasing the Au loading from 0.025 to 0.5 wt. %. However, the further increase in the Au loading caused a decrease in k. These results can be addressed to the loss of photocatalytic active sites on the surface of TiO2 nanoparticles. Based on the TEM images shown in Figure 1, when the Au loading was very high both the amount and the size of Au nanoparticles over the surface of TiO2 nanoparticles were larger, which contributed to a reduction of the adsorption and probably to mitigate the radiation absorbance by the catalytic nanoparticles. Together, these limitations contributed to reducing the photocatalytic efficiency of the nanocomposite towards the samples with lower amounts of Au and demonstrated the relevance of optimizing the Au loading.

**Figure 5.** Photocatalytic degradation of ciprofloxacin (5 mg/L) with bare TiO2 and Au/TiO2 nanocomposite with different Au concentrations under 30 and 180 minutes of UV (**a**) and visible (**b**) radiation. The degradation with bare TiO2 and Au/TiO2 nanocomposites synthesized at different temperatures and Au loading of 0.05 wt. % under 30 and 180 minutes of UV (**c**) and simulated visible light radiation (**d**), respectively.

**Table 1.** Apparent reaction rates (k) for photocatalytic degradation of ciprofloxacin (CIP) (5 mg/L) with bare TiO2 and Au/TiO2 nanocomposite with different Au loadings, over 30 and 180 minutes of UV and simulated visible radiation, respectively.


The photocatalytic assays performed under visible illumination are shown in Figure 6b. Regarding these assays, it is essential first to mention the controls (Supplementary Material, Figures S3 and S4), which have shown that the CIP solution was stable under simulated visible radiation, demonstrating its photostability. Moreover, another control was performed by adding the Au/TiO2 nanocomposites to CIP solution in the dark for 180 minutes. In this case, approximately 11% of CIP was removed from the solution by adsorption to the Au/TiO2 nanocomposites.

**Figure 6.** Degradation efficiency (%) (**a**) and ln (C/C0) vs. time (**b**) for different initial ciprofloxacin concentrations (5, 10, and 25 mg/L), using Au/TiO2 nanocomposites produced at 60 ◦C and with an Au loading of 0.05 wt. %, under 3 hours of simulated visible radiation. Photocatalytic degradation of ciprofloxacin (5 mg/L) in 45 mL of aqueous solution with different Au/TiO2 concentrations (0.1, 0.3, 1.0, and 1.3 g/L). The Au/TiO2 nanocomposite synthesized at 60 ◦C and with an Au loading of 0.05 wt. % was used. The tests were performed over 30 minutes under UV irradiation (**c**).

With the information from controls, it is possible to understand the photocatalytic efficiency of the tested materials better. Similarly, to the UV light experiments, the degradation rates of all produced nanocomposites were faster than that with the bare TiO2. TiO2 could remove ≈ 33% of CIP after 180 min of simulated visible irradiation. This CIP removal may be assigned to adsorption, confirmed by controls performed in the dark (as above mentioned). Additionally, the sun simulator device had a small percentage (≈ 3%) of UV radiation (to mimic sunlight radiation). This radiation can induce a low photocatalytic activity on bare TiO2, which, together with the adsorption of CIP, is responsible for its removal from the solution.

More importantly, the decoration of gold particles on the TiO2 surface resulted in the faster degradation rate of CIP under visible radiation. The bandgap of the composites was lowered, from 2.96 eV to 2.84 eV (Section 2.1). Similar results were obtained for methylene blue degradation using Au/TiO2 nanoparticles. The authors obtained higher degradation efficiencies and ability to use visible radiation [37]. Thus, the materials could absorb the longer wavelength in the visible range (up to 437 nm). The reaction rates' constant increased from 0.073 h<sup>−</sup>1, without Au, to 0.195−0.224 h<sup>−</sup>1, with different Au loadings (Table 1).

The obtained results, for UV and visible radiation, confirmed that the photocatalytic efficiency of the TiO2 nanoparticles was enhanced with the Au loading, until a specific plateau. When the Au loading was higher than 0.1 and 0.05 wt. %, respectively, for UV and visible radiation, the gold nanoparticles can block the surface-active sites of TiO2 nanoparticles [53,54]. Furthermore, an excessive amount of Au nanoparticles can play as recombination centers for photo-induced electrons and holes. Both situations can contribute to a significant reduction of pollutant adsorption and, consequently, the photocatalytic efficiency [55]. The remaining assays of this study will be performed with an Au loading of 0.05 wt. %.

Another critical parameter that is worth to stress and study is temperature, which can affect the surface charge phenomenon and the dispersity of the TiO2 particles in the solution during the synthesis. It can also influence the nucleation and the gold particles' crystal growth on the TiO2 nanoparticle surface. In this study, the synthesis was operated at 25, 60, and 80 ◦C, and the photocatalytic degradation of CIP, with the nanocomposites produced at different temperatures, was performed under UV and visible radiation (Figure 5c,d and Table 2). Regardless of the synthesis temperature, the photocatalytic activity of the nanocomposites (Au loading = 0.05 wt. %) was equal or higher than that of the bare TiO2. Here, the synthesis at the room and medium temperatures (25 and 60 ◦C) yielded the more efficient photocatalytic materials, for UV and visible radiation, towards higher temperature synthesis (80 ◦C).


**Table 2.** Apparent reaction rates (k) for photocatalytic degradation of CIP (5 mg/L) with bare TiO2 and Au/TiO2 nanocomposite synthesized at different temperatures, over 30 and 180 minutes of UV and simulated visible radiation, respectively. The Au loading of 0.05 wt. % was used for the tested materials.

For both types of radiation, the sample obtained at 60 ◦C presented higher degradation efficiencies (Table 2), 91% and 49% of CIP degradation under UV (30 min) and visible radiation (180 min), respectively. On the other hand, the synthesis performed at 80 ◦C revealed lower degradations rates of 80% and 40% for UV and visible radiation, respectively. Other works have reported that higher temperatures accelerate the reduction process and yield broader Au nanoparticles size distributions [56].

In this context, when the synthesis occurred at 80 ◦C, the size of Au nanoparticles produced was larger than the sizes obtained with 25 ◦C and 60 ◦C (in good agreement with STEM-HAADF micrographs, Figure 1). Similarly to what happened with the Au loading, when the amount of Au on the surface of TiO2 was too high, the active sites were blocked and the pollutant adsorption can be limited. Compared with bare TiO2, these results corresponded to a degradation efficiency increase of approximately 13% and 145% for UV and visible radiation, respectively.

Both under UV and visible radiation, another control was performed (Figure 5c,d) by testing single Au nanoparticles at the very same amount of Au (corresponding to 0.05 wt. % obtained at 60 ◦C) and TiO2 nanoparticles on CIP degradation. The results confirmed that the photocatalytic efficiency obtained by the nanocomposites should be assigned to the interface between Au and the TiO2 surface.

#### *2.3. Photocatalytic Degradation*

The rate of photocatalytic degradation depends on the availability of the catalyst surface for the photo-generation of electron-hole pairs that produce hydroxyl radicals. Thus, in these experiments, the amount of catalyst was kept constant, and the number of hydroxyl radicals generated remained the same, while CIP concentration increased. The influence of CIP initial concentration of 5, 10, and 25 mg/L was studied under visible irradiation. It was observed that the CIP concentration impacted by the degradation rate and efficiency (Figure 6). With the lowest CIP concentration, 40% of CIP degradation was obtained after 30 min. With the increase of concentrations by 2 and 5 times, the efficiencies achieved were 22% and 8%, respectively. In these tests, while using the photocatalyst concentration of 0.3 g/L, the adsorption of the CIP on the Au/TiO2 nanoparticles surface might be halted due to surface saturation. Additionally, the presence of organic compounds such as CIP can generate an increased number of intermediates and products, which will compete with CIP for adsorption on the photocatalyst surface [57]. This competition caused a lower reaction rate for high CIP concentration. The following assays, focused on the photocatalytic activity of the produced nanocomposites, were performed using the lowest CIP concentration, 5 mg/L.

In short, the ratio between hydroxyl radical/CIP molecules decreased with higher concentrations, causing lower photocatalytic activity. Moreover, higher CIP concentrations may also reduce radiation harvesting by TiO2 nanoparticles surface, which will also contribute to decreasing the number of hydroxyl radicals formed. Figure 6b displays the plot of ln (C/C0) vs. time at different initial CIP concentrations. Linear plots were observed, and the R<sup>2</sup> values were higher than 0.9, confirming that the photocatalytic degradation of CIP obeyed pseudo first-order kinetics.

The optimal photocatalyst concentration was assessed through degradation of CIP with the different amounts of Au/TiO2 nanocomposites, from 5 to 60 mg, which corresponded to a photocatalyst concentration of 0.1 and 1.3 g/L, respectively. Experimental results are shown in Figure 6c.

In general, with the photocatalytic concentrations of 0.1–1.0 g/L, ≈ 90% of the CIP in solution was degraded, while with the higher concentration of 1.3 g/L, only 50% CIP was degraded after exposure to the same UV irradiation time. At the lowest photocatalytic concentrations of 0.1–0.3 g/L, the photocatalytic degradation increased significantly with the amount of nanocomposite. Indeed, the increased amount of photocatalyst resulted in higher surface coverage, owing to the highest number of active sites [58]. However, when increasing the concentration to 1.0 g/L, the degradation rate remained unchanged. With the highest concentration of Au/TiO2, 1.3 g/L, the degradation was slowest likely because of excessive turbidity that induced light extinguishment after penetrating a short distance from the illuminated surface [59]. Most of the light might be extinguished after penetrating a short distance from the illuminated surface of the suspension. Thus, photocatalytic particles in an inner region could not be activated. The result agrees with other reports [58,60]. For this reason, the concentration of 0.3 g/L (15 mg in 45 mL), which yielded the highest photocatalytic efficiency, was chosen for the following photocatalytic activity assays.

The reproducibility of the nanocomposites efficiency was also tested using three independent syntheses, performed under the same conditions. The produced samples were then used in the photocatalytic degradation of CIP under visible radiation (results in the Supplementary Material, Figure S5). The apparent reaction rate constant of the three experiments fluctuated around the value of 0.219 <sup>±</sup> 0.022 min−1. The standard deviation of 10% proved the reproducibility of the method to enhance the photocatalytic activity of pure TiO2 nanoparticles and endow them with visible light activity. It is also important to clarify that the Au concentrations used in this study were nominal, as no inductively coupled plasma atomic emission spectroscopy (ICP-AES), or similar characterization, was performed. However, given the considerable reproducibility of the method, the putative loss of gold would by similar for all the Au concentrations tested, making them comparable.

#### **3. Computational Modeling: Gold on Titanium Dioxide and Charge Transfer**

A computational study was performed to rationalize the effect of gold nanoparticles on TiO2. The GFN-xTB (Geometry, Frequency, Noncovalent, eXtended Tight-Binding) was used. GFN-xTB is a new semiempirical method developed by Grimme et al. [61] that allows computing efficiently systems with thousands of atoms. The GFN-xTB software used (version 5.4.6) did not allow us to compute systems with periodic boundary conditions, so a finite system composed of a gold nanoparticle adsorbed on a larger TiO2 nanoparticle was used. We chose a cuboctahedral (TiO2)97 anatase nanoparticle, which was found to produce bulk-like electronic properties [62] and had two large, equal, flat surfaces of ~ 1.2 <sup>×</sup> 1.5 nm<sup>2</sup> on which a cuboctahedral (Au55) <sup>−</sup>3, 10 nm diameter, gold nanoparticle was adsorbed [63]. Several adsorption modes were possible, yet an exhaustive search of these was beyond the scope of this study. We positioned the gold nanoparticle on four, arbitrary, different orientations so that, in all cases, one of the faces of (Au55) <sup>−</sup><sup>3</sup> was parallel to the anatase flat surface and minimized. In this minimization, the TiO2 coordinates were held fixed while the Au first neighbors' distances were constrained harmonically to an equilibrium value to force the gold nanoparticle to keep its initial shape while retaining some flexibility. Two different pH conditions were considered: (1) A neutral/basic pH represented by the bare (TiO2)97 anatase nanoparticles (i.e., without protonation), and (2) an acidic pH, where the eight under-coordinated oxygen atoms in the anatase nanoparticle were protonated, (TiO2)97H8 <sup>+</sup>8, which should represent better the experimental conditions (Figure 7).

**Figure 7.** The (Au55(TiO2)97H8) <sup>+</sup><sup>5</sup> with standard atom coloring (left) with color-rendered atomic charges (right), so that red represents negatively charged atoms, blue positively charged atoms, and white neutral atoms.

To study the charge transfer between (Au55) <sup>−</sup><sup>3</sup> and (TiO2)97 and ((TiO2)97H8) <sup>+</sup>8, we computed atomic charges specifically suited for condensed phases [64]. The analysis of the charges in (Au55(TiO2)97) <sup>−</sup><sup>3</sup> and (Au55(TiO2)97H8) <sup>+</sup><sup>5</sup> (Figure 7) show that the Au nanoparticle transferred electrons to the anatase, namely 4.7 electrons for neutral TiO2 and 4.3 electrons for the protonated TiO2. Most of this charge was transferred directly to Ti atoms, namely 3 and 3.3 electrons for the neutral and protonated case, respectively. So, the Au acted as an electron donor populating the Ti(d) states that were responsible for the photocatalytic activity of the material. Consequently, Au may increase the catalytic activity of the composite material through interfacial electron transfer. Interestingly, the anatase surface also polarized the gold nanoparticle so that all atoms in direct contact with TiO2 were more oxidized, i.e., had larger positive charges, see Figures 7 and 8.

This effect was observed in all cases, for neutral and acidic pH and when the harmonic constraint on Au atoms was released, which indicates that the observed charge transfer is a fundamental process of the Au-TiO2 interface. This corroborated the experimental controls shown in Figure 5c,d, showing that separation of Au and TiO2 nanoparticles yielded lower efficiencies. Moreover, this finding also fit previous DFT (density functional theory) calculations which found that an Au nanorod on a rutile TiO2 (110) surface might act as an activator for molecular oxygen through charge transfer to nearby Ti+<sup>4</sup> atoms [65]. Mechanistically, the presence of Au activated superficial Ti+<sup>4</sup> atoms nearby for catalysis via direct charge transfer, which rationalized our experimental observation that lower Au loading and small gold nanoparticles had larger catalytic activity than larger loadings and nanoparticles on TiO2.

**Figure 8.** The (Au55(TiO2)97H8) <sup>+</sup><sup>5</sup> with only the titanium subnetwork and gold atoms with their corresponding point charges.

#### *Adsorption of Ciprofloxacin on Au55(TiO2)97H8* +*58*

We also explored the different energetics of a CIP molecule interacting with (Au55(TiO2)97H8) +58 on four different adsorption sites (Figure 9). All the adsorption modes were found to be binding, i.e., exothermic, with energies ranging between 0.7 and 2.4 eV. The strongest binding was observed for the structure with a large contact between the oxygen atom of the carboxylic group and the gold nanoparticle, 2.4 eV. This binding was similar in energy to the adsorption on the anatase clean surface (2.1 eV), which indicates that both processes might be competitive and that CIP molecules might also adsorb near the gold/anatase interface where the Ti atoms were activated through electron transfer.

**Figure 9.** Four possible adsorption geometries of CIP on (Au55(TiO2)97H8) <sup>+</sup><sup>58</sup> with corresponding binding energies. The positive binding energies indicate that the process is exothermic.

#### **4. Materials and Methods**

#### *4.1. Materials*

P25 TiO2 nanoparticles were kindly provided by Evonik (Essen, Alemanha). Gold(III) chloride trihydrate, 99.9% CAS: 16961-25-4 (liquid solution) was purchased from Sigma-Aldrich (St. Louis, Missouri, EUA). Sodium hydroxide (NaOH) was obtained from VWR (Radnor, Pensilvânia, EUA) Millipore Milli-Q-system ultra-pure (UP) water was used in all the experiments.

#### *4.2. Nanocomposite Synthesis*

The Au/TiO2 nanocomposites were synthesized, as illustrated in Figure 10, dispersing 200 mg of TiO2-P25 nanoparticles in 40 mL of ultra-pure (UP) water in a sonication bath for 30 min. Afterwards, this solution was placed under agitation in a water bath at different temperatures (25, 60, and 80 ◦C), using a thermostat to precisely control and stabilize the temperature, avoiding thermal gradients. When the dispersion solution reached the desired temperature, different volumes from the chloroauric solution (10 μL of Gold(III) chloride trihydrate in 100 mL of UP water) were added to achieve the Au loadings of 0.025, 0.05. 0.1, 0.25, and 0.5 wt. %. The solution was then stirred for 10 minutes to achieve a homogeneous distribution of gold precursor solution. Later, several volumes of a 0.1 M sodium hydroxide solution (NaOH) were added dropwise and mixed for 10 minutes to obtain a pH = 9. The solution was then centrifuged at 23,000 rpm, the supernatant discarded, and the nanocomposite pellet redispersed in UP water with the ultrasonication for 1 minute, and this washing procedure was repeated one more time. The last step was to dry the nanocomposite at 80 ◦C in an oven overnight and grind it with a pestle and mortar to obtain a fine powder.

**Figure 10.** Schematic representation of the main steps to synthesize Au/TiO2 nanocomposites trough nanoprecipitation.

#### *4.3. Characterization*

The morphology of the nanocomposites was assessed by transmission electron microscopy (TEM), a Tecnai T20 from FEI (Hillsboro, Oregon, EUA). For the analysis, the nanocomposite samples were sonicated for 5 minutes to achieve good dispersion and afterwards a drop of the suspension was placed on a copper grid and dried at room temperature for the analysis. Particle size histograms were obtained after measuring at least 200 nanoparticles using Image J 1.50i software. Aberration-corrected scanning transmission electron microscopy (Cs-corrected STEM) images were acquired using a high-angle annular dark field detector in an FEI XFEG TITAN (Hillsboro, Oregon, EUA) electron microscope operated at 300 kV equipped with a Spherical Aberration Corrector for Transmission Electron Microscopes (CETCOR) Cs-probe corrector from CEOS Company (Heidelberg, Germany), allowing the formation of an electron probe of 0.08 nm. Elemental analysis was carried out with an EDX (energy-dispersive X-ray spectroscopy) detector, which allows performing EDX experiments in the scanning mode.

The crystallographic phases of the pure TiO2 and the Au/TiO2 nanocomposite were evaluated by X-ray diffraction using a D8 Discover diffractometer with incident Cu Kα (40 kV and 30 mA), from Bruker (Billerica, Massachusetts, EUA).

The average hydrodynamic diameter was assessed by dynamic light scattering (DLS) in a Zetasizer NANO ZS-ZEN3600, Malvern (Malvern Instruments Limited, United Kingdom), equipped with a He–Ne laser (wavelength 633 nm) and backscatter detection (173◦). The samples were dispersed (0.1 mg/L) in ultrasonication bath at 22 ◦C for 30 minutes to avoid aggregates, and each sample was measured 10 times. The zeta (ζ) potential was measured in the same device, and TiO2 nanoparticles were equally suspended in ultra-pure water and solutions at different pHs (2, 4, 7, 9, and 12) were prepared with HCl (1M) and NaOH (1M) solutions. The results were obtained using the Smoluchowski theory approximation, and each sample was measured 10 times at 22 ◦C. The manufacturer software (Zetasizer 7.12) was used to assess particles diameter (intensity distribution), the polydispersity index (PDI), and z-potential values.

The optical properties of the pristine TiO2 and the Au/TiO2 nanocomposite were assessed by UV–vis reflectance, using a Shimadzu UV-2501-PC (Kyoto, Japan) equipped with an integrating sphere. The spectra were acquired in reflectance, and the bandgap was estimated via the Kubelka–Munk Equation (1) [52] and the Tauc plot represented by Equation (2).

$$F(\mathbb{R}) = (1 - R\_{\infty})^2 / (2R\_{\infty}) \tag{1}$$

where *R*<sup>∞</sup> (*R*Sample/*R*BaSO4) corresponds to the reflectance of the sample and *F*(*R*) is the absorbance.

$$\left[F(R)h\nu\right]^{1/n}\text{ versus }h\nu\tag{2}$$

where *h* is the Planck constant (6.626 <sup>×</sup> 10−<sup>19</sup> J), *n* is the frequency, and *n* is the sample transition (indirect transition, *n* = 2) [66].

#### *4.4. Photocatalytic Degradation*

The photocatalytic activity of all the produced samples and pristine TiO2 was assessed by performing (CIP) degradation tests, under artificial ultraviolet (UV) or visible illumination. First, a solution of 5 mg/L of CIP was prepared. The CIP solution was adjusted to pH = 3, to ensure the solubility, by using 0.1 mL hydrochloric acid (HCl) 1 M.

Before the degradation assays (UV or visible radiation), the Au/TiO2 or P25 nanoparticles were stirred in the dark for 30 min to achieve an adsorption-desorption equilibrium.

The UV degradation of CIP was performed in a chamber with six Philips 8 W mercury fluorescent lamps with the mode wavelength of 365 nm. The suspensions of photocatalysts and CIP were kept stirred in a container under the illumination from the top. The distance between the beaker and the lamp was 13.5 cm, and the intensity coming to the system was 15−17 W/m2. The samples were irradiated for 30 min.

The visible light tests were performed in a visible chamber fabricated by Ingenieurbüro Mencke & Tegtmeyer GmbH©, Hameln, Alemanha. According to the manufacturer, the visible light spectrum was equivalent to that of the natural solar light. The light source had an intensity of 98 W/m2. The visible light test was performed similarly to the UV test. Here, the container was placed at 21 cm from the light source, and the samples were irradiated continuously for 180 minutes.

The first photocatalytic activity tests were performed to determine the optimal ratio of CIP/catalyst. For this purpose, 5, 15, 45, and 60 mg of Au/TiO2 nanocomposite were dispersed in a borosilicate beaker of 80 mL with 45 mL of CIP solution (5 mg/L). The effect of Au loading on the photocatalytic efficiency was also assessed, under UV and simulated visible radiation. The impact of the synthesis temperature (25, 60, and 80 ◦C) on CIP photocatalytic degradation was equally evaluated using both types of illumination. The photocatalytic reproducibility tests were performed using nanocomposites produced in different batches but under the same synthesis conditions.

The bare TiO2 nanoparticles were used as controls in all the assays. Additionally, to prove the relevance of the Au/TiO2 nanocomposites' structure and interface, the photocatalytic degradation of CIP was assessed using the same amounts of Au and TiO2 nanoparticles, not as a nanocomposite, but separately added to the solution.

The photocatalytic efficiencies were tested by degrading CIP in aqueous solution under UV and visible radiation and monitoring the maximum absorption peak (277 nm) using a Shimadzu UV-2501PC UV/Vis spectrophotometer. The degradation fit the Langmuir-Hinshelwood model, expressed by Equation (3):

$$\text{C}/\text{C}\_{0} = \exp^{-kt} \tag{3}$$

where *C*<sup>0</sup> and *C* represent the concentration of the pollutant at time 0 min and at time *t*, respectively, and *k* is the first-order rate constant of the reaction.

#### **5. Conclusions**

An Au/TiO2 nanocomposite was produced, characterized, and applied in the photocatalytic degradation of ciprofloxacin (CIP). The characterization results changing the synthesis conditions (temperature and Au loading) indicated that the synthesis performed at 60 ◦C with the Au loading of 0.05 wt. % yielded the most homogeneous distribution of Au nanoparticles (≈3 nm) over TiO2 nanoparticles surface, after TEM inspection. Additionally, these samples absorbed more radiation in the visible range (≈66% at 545 nm) and presented a lower bandgap (2.84 eV vs. 2.96 eV from bare TiO2). The photocatalytic results confirmed that all the manufactured nanocomposites possessed higher photocatalytic efficiency in the UV and simulated visible radiation towards the pristine TiO2. It was also possible to understand the impact of the synthesis parameters envisaging the optimal photocatalytic efficiency conditions. In this way, with the Au/TiO2 nanocomposite, it was possible to enhance the photocatalytic degradation efficiency in 13% and 145% under UV and simulated visible light radiation, respectively. The gold nanoislands played a paramount role transferring electrons from Au to the anatase from TiO2 nanoparticles. Additionally, Au endowed the nanocomposite with the ability to absorb the visible radiation.

Computational modeling supported the experimental data, showing the ability of Au to bind TiO2 anatase surfaces as well as the relevant role of Au transferring electrons. The fundamental importance of the interface between TiO2 and Au nanoparticles regarding the enhanced photocatalytic activity was also rationalized. Moreover, models indicated a high affinity of CIP to both Au and TiO2 surfaces, which favors the adsorption process and consequently may also be cause for enhanced photocatalytic efficiency in the presence of Au nanoparticles.

According to the results obtained through systematic experimental data and modeling results, the simple method herein presented constitutes a reliable approach to produce efficient photocatalytic materials.

**Supplementary Materials:** The following are available online at http://www.mdpi.com/2073-4344/10/2/234/s1: Figure S1: Size distribution of Au nanoparticles for synthesization at 25 and 80 ◦C; Figure S2: Photostability of CIP solution under UV; Figure S3: Photostability of CIP solution under visible radiation; Figure S4: Synthesis reproducibility on CIP degradation.

**Author Contributions:** Conceptualization, P.M. and S.L.-M.; data curation, P.M., S.K., and H.N.L.; formal analysis, H.N.L., L.P., M.M.-F., and S.L.-M.; investigation, P.M., S.K., H.N.L., and M.M.-F.; methodology, P.M., S.K., and V.S.; project administration, S.L.-M.; resources, V.S., M.A., G.C., and S.L.-M.; software, M.M.-F.; supervision, K.K., M.A., G.C., and S.L.-M.; validation, V.S.; visualization, P.M. and M.M.-F.; writing—original draft, P.M. and H.N.L.; writing—review & editing, P.M., M.A., and S.L.-M. All authors have read and agreed to the published version of the manuscript.

**Funding:** The authors acknowledge funding from the Basque Government Industry Department under the ELKARTEK Program and the Spanish Ministry of Economy and Competitiveness (MINECO) through the project MAT2016-76039-C4-3-R (AEI/FEDER, UE) (including the FEDER financial support). This work was also supported by the Graduate Academy of the Technische Universität Dresden. Centro de Investigacion Biomédica en Red – Bioengenharía, Biomateriales e Nanomedicina (CIBER-BBN) is an initiative funded by the 6th National R&D&i Plan 2008–2011, Iniciativa Ingenio 2010, Consolider Program, and CIBER Actions and financed by the Instituto de Salud Carlos III (Spain) with assistance from the European Regional Development Fund. S. Kappert and H.N. Le acknowledge fruitful discussions with Nadia Licciardello.

**Acknowledgments:** This work was supported by the Portuguese Foundation for Science and Technology (FCT) in the framework of the strategic projects UID/FIS/04650/2013 by Fundo Europeu de Desenvolvimento Regional (FEDER) funds through the COMPETE 2020—Programa Operacional Competitividade e Internacionalização (POCI) with the reference project POCI-01-0145-FEDER-006941, project PTDC/CTM-ENE/5387/2014, as well as UID/BIO/04469 unit through COMPETE 2020 (POCI-01-0145-FEDER-006684) and BioTecNorte operation (NORTE-01-0145-FEDER-000004) funded by the European Regional Development Fund under the scope of Norte2020—Programa Operacional Regional do Norte. P.M. Martins thanks the FCT for the grant SFRH/BD/98616/2013 and Luciana Pereira for the grant SFRH/BPD/110235/2015. M. Melle-Franco would like to acknowledge support from Centro de Investigação em Materiais Cerâmicos e Compósitos (CICECO)—Aveiro Institute of Materials, POCI-01-0145-FEDER007679 (UID/CTM/50011/2013) and the FCT (IF/00894/2015).

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


© 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

### *Article* **TiO2 and Active Coated Glass Photodegradation of Ibuprofen**

**Samer Khalaf 1,2,\*, Jawad H. Shoqeir 1, Filomena Lelario 2, Sabino A. Bufo 2, Rafik Karaman 2,3 and Laura Scrano <sup>4</sup>**


Received: 29 January 2020; Accepted: 25 February 2020; Published: 18 May 2020

**Abstract:** Commercial non-steroidal anti-inflammatory drugs (NSAIDs) are considered as toxic to the environment since they induce side effects when consumed by humans or aquatic life. Ibuprofen is a member of the NSAID family and is widely used as an anti-inflammatory and painkiller agent. Photolysis is a potentially important method of degradation for several emerging contaminants, and individual compounds can undergo photolysis to various degrees, depending on their chemical structure. The efficiency oftitanium dioxide (TiO2) and photocatalysis was investigated for the removal of ibuprofen from the aquatic environment, and the performance of these different processes was evaluated. In heterogeneous photocatalysis, two experiments were carried out using TiO2 as (i) dispersed powder, and (ii) TiO2 immobilized on the active surface of commercial coated glass. The kinetics of each photoreaction was determined, and the identification of the photoproducts was carried out by liquid chromatography coupled with Fourier-transform ion cyclotron resonance mass spectrometry (LC-FTICR MS). The overall results suggest that the TiO2 active thin layer immobilized on the glass substrate can avoid recovery problems related to the use of TiO2 powder in heterogeneous photocatalysis and may be a promising tool toward protecting the environment from emerging contaminants such as ibuprofen and its derivatives.

**Keywords:** ibuprofen; advanced oxidation process; TiO2; photocatalysis; active glass

#### **1. Introduction**

Emerging contaminants resulting from the presence and circulation of pharmaceuticals (PhCs) were the focus of many environmental chemists over the last few decades. In the aquatic environment, PhCs are introduced anthropogenically through pharmaceutical or conventional plants [1]. PhCsare found in tiny concentrations in surface waters, indicating insufficient treatment of such entities during the standard sewage treatment processes (STPs). The occurrence of these toxic drugsin wastewater effluent, along with their metabolites which may be much more harmful than their parent compounds [2,3], has the potential to be a great health problem since they are endocrine-disrupting agents, thus posing a significant barrier to the use of water recycling [4].

Ibuprofen (IBP), (*RS*)-2-(4-(2-methylpropyl) phenyl) propanoic acid, shown in Figure 1, is a non-steroidal anti-inflammatory drug(NSAID) belonging to the class of propanoic acid derivatives used as pain relief for several inflammation conditions, including rheumatoid arthritis, asan analgesic for pain relief in general, and as an antipyretic to help in fever conditions [5]. IBP enters the aquatic environment through effluents exiting secondary wastewater treatment plants, which are inefficient in removing a variety of small organic molecules, particularly pharmaceuticals [6]. Reported studies showed that the concentrations of IBP found in rivers and other environmental waters range between 10 ng·L−<sup>1</sup> and 169 <sup>μ</sup>g·L−<sup>1</sup> [7].

**Figure 1.** Chemical structure and ultraviolet (UV) absorbance of ibuprofen (IBP).

The fact that the current conventional wastewater treatment technologies such as those based on biological, thermal, and physical treatment processes are not efficient in removing or degrading small-molecular-weight pharmaceuticals with low biodegradability and high chemical stability such as ibuprofen [8] encouraged us to devote considerable effort toward developing a novel purification method that can efficiently remove this recalcitrant organic contaminant from the water environment.

Recently, we found that the integration of separation technologies, consisting of sequential elements of ultra-filtration (UF), activated carbon filtration (AC), and reverse osmosis (RO), as well as adsorption technology based on a surface modified clay minerals, was efficient in removing IBP and other pharmaceuticals to a safe level [9–13]. Nevertheless, the operating principles of these tools are only based on phase-transfer technologies, whereby the contaminant is retained on the filter or adsorbent without being degraded or destroyed to non-toxic compounds. Furthermore, some of the technologies used, such as UF and RO, are too expensive to be adopted in most real environmental situations. For the abovementioned reasons, we successfully attempted to find a good alternative method (degradation via photocatalysis) to these technologies for removing such pollutants from the aquatic environment.

The growing awareness of the risk arising from the occurrence of toxic organic contaminants in the aquatic environment promoted the development of technologies, such as photodegradation, and other advanced oxidation processes (AOPs), for efficient destruction of organic toxic compounds that exist in water and wastewater, including PhCs [14–16].

AOPs, i.e., processes based on highly reactive species such as hydroxyl radical (•OH), can oxidize and mineralize practically every organic entity [3], yielding CO2 and inorganic ions, thus resulting in total destruction of the target pollutant [3,8,16]. Advanced oxidation processes (AOPs) involve several homogeneous and heterogeneous processes such as photolysis, photocatalysis, ozonation, electrochemical oxidation, photo-Fenton, wet air oxidation, and sonolysis [8,15,16]. The most popular and effective type of AOP employed in water and wastewater treatment is heterogeneous photocatalysis with semiconductors [17–19].

Heterogeneous photocatalysis is a process via which a photoreaction is accelerated by the presence of a catalyst (usually semiconductor). In order for this to occur, the dispersed solid particles of the semiconductor in the treated solutionshould absorb significant portions of the UV light, and, under radiation, they may be photo-excited and produce oxidizing agents from water and oxygen [19]. Generally, TiO2 is considered the most efficient semiconductor to be employed in photocatalysis because of several factors including its low cost, low toxicity, chemical stability, large band gap, and high photosensitivity [8,16–19].

Under ultraviolet irradiation, TiO2 as a semiconductor causes the jump of an electron (e−) from the valence band (VB) to the conduction band (CB), resulting in the formation of a positive hole (h+) at the site of the electron. In the presence of aqueous suspended TiO2, the hole and electron can produce radicals of hydroxyl and superoxide that are very potent in the oxidation of many kinds of organic entities found in water sources, thereby leading to total degradation of these toxic organic agents. [20,21]. Equations (1)–(3) depict the formation reactions of the superoxide and hydroxyl radicals upon catalysis with TiO2

$$\rm{TiO}\_2 + \rm{hv} \rightarrow \rm{e}\_{\rm{CB-}} + \rm{h}\_{\rm{VB+}}.\tag{1}$$

$$\rm H\_2O + h\_{VB+} \rightarrow OH^\bullet + H\_{aq+} \tag{2}$$

$$\rm O\_{2(ads)} + e\_{CB\cdot} \rightarrow O\_2{\bullet}^\bullet \text{-(ads)} + H\_2O + OH^\bullet. \tag{3}$$

In some processes, a complete degradation of organic pollutants requires the presence of a radiation source, oxidizing agent, and a semiconductor. Oxidation and reduction processes are promoted by photo-generated charge carriers resulting from the excitation of TiO2 via photons with higher energy. Currently, this kind of photocatalysis is utilized to purify water [8,16–19].

Although photocatalytic degradation using suspensions of TiO2 particles was extensively employed to catalyze different contaminants, such as drugs, and although it achieved good results in recent years, this technique fails to be widely used because of the high cost and difficulty in isolating the semiconductor from the mixture after degradation [8,16–19]. To find a way around the need for catalyst recovery via filtration, a different approach, consisting of catalyst immobilization on a stationary support, should be assessed. For fulfilling this aim, TiO2 immobilized on different materials instead of the traditional powder was advocated and tested for obtaining a promising clean treatment method with a low cost [22–24].

In this study, the efficiency of two different systems, direct photolysis and heterogeneous photocatalysis (TiO2 powder and TiO2 immobilized on active glass), was investigated using simulated solar irradiation for the removal of IBP and its major photoproducts from the aqueous phase. The comparative performance of the adopted process was analyzed under the same experimental conditions, and the kinetics for each photodegradation reaction was evaluated. Moreover, major photoproducts were detected and identified via liquid chromatography coupled with Fourier-transform ion cyclotron resonance mass spectrometry (LC-FTICR MS).

It should be emphasized that the study on the photodegradation of IBP using an immobilized TiO2 system can be regarded as representative of a process for the degradation of a variety of pollutants which impose a risk to the environment.

#### **2. Results and Discussion**

#### *2.1. Characterization of Active TiO2-Coated Glass*

Figure 2A,B depict the SEM image of the active glass and the TiO2 coating comb geometry on the glass surface, respectively. The TiO2 film thickness was 397.2 nm, as shown in Figure 2A.

**Figure 2.** (**A**) SEM image of the blue glass cross-section illustrating the position of the TiO2 layer immobilized on the glass surface; (**B**) SEM image of the blue glasssurface illustrating the fine-tooth comb nature of the TiO2 coating.

Table 1 lists the elements in the glass surface and core class. As shown in the table, TiO2is present only in the glass surface along with other elements, whereas it is absent in the core glass.


**Table 1.** Scanning electron microscopy EDX analysis of the glass surface coated with TiO2 as compared to the glass core composition.

Sheel et al. (1998) reported the presence of cobalt oxide in a concentration of less than 75 μg/g [24], whereas we did not succeed in detecting any amount of cobalt oxide using our technique. It is believed that the blue color of the glass is due to the presence of cobalt oxide. For this reason, we report here the TiO2-coated active glass using the abbreviation "blue glass".

The glass coating is based on photocatalytic anatase TiO2, which is the most effective known photocatalyst [22,23]. For the preparation of the glass coating, TiO2 nanocrystallinefilm is deposed onto float glass using an atmospheric pressure chemical vapor deposition technique (APCVD) as described in Reference [24]. Anatase is considered a more efficient photocatalyst than rutile because of its slower rate of recombination [23,24]. In the coated surface of the active glass, the catalyzing efficiency is improved by the presence of Fe3O4, which favors the formation of multiple band gaps, enlarging the wavelength range that can be absorbed by the glass surface [22,23].

#### *2.2. Photodegradation Experiments*

#### 2.2.1. Preliminary Study

Prior to photodegradation of IBP, which involves photolysis and photocatalysis experiments, thermal (45 ◦C) or hydrolytic reactions in pure IBP solution and the adsorption of the pharmaceutical on the catalyst surface in TiO2 powder suspension were assessed. No significant loss of IBP occurred in dark conditions due to thermal reactions or hydrolysis. The adsorption equilibrium on TiO2 powder was reached within 60 min, and a slight decrease (not more than 4.3%) of free IBP concentration in the 0.2–0.3 g·L−<sup>1</sup> TiO2 suspensions was achieved. Adsorption increased to 6.9% and more (7.8%) when the amount of TiO2 powder was augmented to 0.4–0.5 g·L−1. The mentioned values are the average of three replicates.

#### 2.2.2. Photolysis Experiment

The uranyl oxalate method [25,26] was used to assess the light emission effectiveness of the irradiation system prior to the experimental work. The disappearance of oxalate was 7.2 <sup>×</sup> <sup>10</sup>−<sup>4</sup> mol·s–1.

Despite the high efficiency of the irradiation system (irradiance to exposed surface of the reactor, 500 W/m2), IBP concentration during this experiment was decreased only by 8.4% after more than 20 h. This result indicates that IBP is stable during photolysis because it has great chemical stability and a reduced molar adsorption coefficient above 280 nm [19] (Figure 1).

#### 2.2.3. Photocatalysis Experiments

TiO2 commercial powder and TiO2-coated active glass were employed separately. The degradation efficiency via the two different methods was compared.

#### Photocatalytic Degradation Using TiO2Powder

The concentration of IBP during the photocatalysis reaction was monitored using High Performance Liquid Chromatography-Ultra Violet HPLC-UV. The standard solution used showed a peak of IBP at 4.1 min retention time. After 5 min of sample irradiation, a marked decrease in the IBP concentration was observed (21.8% of the initial concentration) along with the appearance of a new photoproduct (Figure 3A,B).

After 30 min, a new peak due to the formation of another derivative was observed, and an approximately 60% reduction in IBP initial concentration was obtained (Figure 3C). Complete disappearance of IBP was achieved after 270 min (Figure 3D), while complete depletion including derivatives was observed after approximately 23hours (Figure 3E).

The combined results demonstrate that IBP was completely degraded by photocatalytic oxidation using TiO2 powder under simulated solar irradiation, and the efficiency of its removal was more than 87% within 80 min. The degradation of IBP occurred as a result of a photo-irradiation of the semiconductor, causing an electron transfer to the conduction band which subsequently formed a hole in the valence band, which led to photo-induced charge separation on the semiconductor surface and an exchange of electrons on the water semiconductor interface. This led to the formation of •O2 − via interactions of adsorbed oxygen molecules with the photo-generated conduction band electrons, whereas the •OH generated from the oxidation of the adsorbed water or hydroxyl anions by the valence band hole oxidized the adsorbed IBP molecules [17,19].

**Figure 3.** HPLC-UV separation of photodegraded solution using TiO2 powder, IBP (1), and IBP photoproducts (2 and 3) (**A**) at time zero (initial standard solution), (**B**) after 5 min, (**C**) after 30 min, (**D**) after 270 min with complete disappearance of IBP, and (**E**) after 23 h with complete disappearance of IBP photoproducts.

Photocatalytic Degradation Using TiO2-Coated Active Glass

During the first 2 h, a slight decrease in IBP concentration (2.7%) was achieved, while the appearance of a unique photoproduct was accomplished after 3 h (Figure 4A,B). Following a further 9hof irradiation, the concentration of IBP was decreased to 50%, and, after 24 h, IBP and its photoproduct disappeared (Figure 4C).

**Figure 4.** HPLC-UV separation of photo-degraded solution using TiO2-coated blue glass, IBP (1), and IBP photoproducts (2) (**A**) after 180 min, (**B**) after 10 h, and (**C**) after 24 h with complete disappearance of IBP and IBP photoproducts.

#### *2.3. Kinetics Studies*

#### 2.3.1. Experimental Observations

No degradation of IBP was observed in the dark in all aqueous environments adopted for the experiments. Direct photolysis under simulated sunlight did not achieve the desired goal. Accordingly, we can conclude that the direct interaction of IBP with sunlight (both via thermal hydrolytic reactions and photolysis) cannot lead to IBP's quick degradation. However, in the presence of TiO2, complete removal of this NSAID was obtained although a xenon lamp with low UV energy was used for irradiation aiming at the simulation of sunlight effect.

Different amounts (0.1–0.5g·L−1) of TiO2 micro-particles were added to a solution of 25 mg·L−<sup>1</sup> IBP to determine the efficiency of the catalytic process. The half-life (experimentally observed) of the mother molecule was reduced upon increasing the concentration of the catalyst from 0.1 to 0.2 g·L<sup>−</sup>1, and it remained constant upon adding an amount of 0.3 g·L<sup>−</sup>1, while it increased when concentrations of 0.4 or 0.5 g·L−<sup>1</sup> were tested. The rationale behind such behavior is that the number of IBP molecules and photons absorbed on the TiO2 particles increased with a moderate increase in the catalyst loading; however, with further addition of the semiconductor (powder), the phenomenon of light scattering took place and the number of useful photons per mass unit of TiO2 was reduced. The disappearance of IBP at the highest concentrations of TiO2 powder was mostly due to its mere physical adsorption onto the surface of semiconductor particles.

Figure 5A,B illustrates the depletion trend of IBP measured as Ct/C0 versus irradiation time (A) and evolution of photoproducts (B) using different photodegradation methods. In both photocatalysis

processes, IBP underwent complete disappearance via the formation of one or two intermediates that were subsequently removed within 24 h.

**Figure 5.** Evaluation of IBP degradation measured as Ct/C0 versus irradiation time (**A**); evolution of photoproducts using different photodegradation methods (**B**).

In the photocatalysis experiment with TiO2-coated active glass, the reaction was apparently slower than degradation obtained using TiO2 powder, but a satisfactory depletion of IBP and its derivatives was reached in approximately the same time.

#### 2.3.2. Kinetic Parameters

To find the kinetics model, kinetic parameters were calculated using integrated equations describing zero-, first-, and second-order (Langmuir-Hinshelwood) order equations [27]. According to Snedecor and Cochran (1989) [28], the least square method should be utilized to find the best fit.

Table 2 summarizes the kinetic parameters of IBP degradation under the photocatalysis experiment with TiO2 powder.


**Table 2.** Kinetic parameters of IBP degradation under photocatalysis experiments.

ΣLSq, sum of least squares = Σn (Cexp–Ccalc)2; Cexp, experimental values of concentrations; Ccalc, value of concentrations calculated from rate equations; n, number of experimental observations; k, kinetic constant; t1/2, half-life.

It must be taken into account that, owing to the dissimilar units associated with them, the values of kinetic constants calculated by equations describing reactions of different order cannot be compared. For this reason, it is useful to consider the values of half-life, which are always expressed in time units. Table 2 shows dissimilar values of the half-life when calculated using different equations applied to the same system. The least square method of estimation is a powerful method to assess the equation that can best fit the experimental data.

Apparently, the measured reaction rate of IBP under irradiation conditions using TiO2 powder as a catalyst was best fit by a Langmuir–Hinshelwood-type equation [29].

$$\mathbf{C}\mathbf{t} = \mathbf{C}\_0 \,\mathbf{t}\_{1/2}/(\mathbf{t} + \mathbf{t}\_{1/2}),\tag{4}$$

where C0 is the initial amount (mg) of IBP per liter of solution, Ct is the remaining concentration at time t, and t1/<sup>2</sup> is the half-life of the reactant.

Equation (4) shows the minimum value of the sum of least squares, based on the number of observations (ΣLSq)/n, and describes a second-order reaction governed by the kinetic law.

$$\mathbf{v} = -\mathbf{d}\mathbf{C}\_t \text{/dt} = \mathbf{k}\mathbf{C}\_t \text{\textdegree } \tag{5}$$

where ν is the reaction rate, and k is the rate (or kinetic) constant [27,29], which in our case can be calculated as

$$\mathbf{k} = 1/(\mathbb{C}\_0 \mathbf{t}\_{1/2}).\tag{6}$$

Equation (5) represents a double dependence of the reaction velocity on the concentration.

The rationale behind such a finding may be due to the total reaction rate during the photocatalytic process being affected by two sorption states, both depending on the dissolved concentration of the pharmaceutical. The amount of reactant disappearing at each time t is affected by its free concentration in the powder suspension, as well as by the amount adsorbed on the catalyst particles, which depends on the remaining free concentration of IBP.

The half-life value for a second-order reaction, calculated by means of the linearized form of Equation (4) (Table 2), was just 11.8 min, while, after 80 min, 87% of IBP was converted.

The second-order kinetics shown in Figure 6A was confirmed by the linear behavior of (C0/Ct) as a function of irradiation time (Figure 6B).

**Figure 6.** (**A**) Photodegradation of ibuprofen catalyzed by TiO2 powder; Ct calc, values calculated using Equation (4); Ct exp, experimental values; error bars represent the standard deviations of three replicate experiments. (**B**) Trend of second-order linearized equation used for the calculation of kinetic parameters reported in Table 2.

From the results, it can be remarked that the initial degradation rate was high; however, it decreased rapidly as the reaction proceeded. The degradation was fast during the first 20 min, and then it gradually decreased; this trend is typical of second-order reactions. Several observations can be related to such a behavior: (1) the high concentration of IBP at the beginning of the reaction facilitates the useful attack by the hydroxyl radicals, resulting in high degradation rate; however, when IBP concentration gradually decreases, the degradation rate subsequently decelerates due to the dilution effect that reduces the possibility of useful collisions with the hydroxyl radicals; (2) the competitive reactions of the hydroxyl radicals with IBP degradation products that are produced during the reaction; (3) the recombination reactions of radical–radical.

The photodegradation reaction of IBP catalyzed by TiO2 immobilized on active glass surface achieved 85% compound disappearance after 24 h of simulated solar light irradiation. By attempting to fit the concentration values vs. time using various-order integrated kinetic equations, it was found that the data best fit the first-order kinetic equation

$$\mathbf{C}\_{\mathrm{I}} = \mathbf{C}\_{\mathrm{O}} \stackrel{-\mathrm{k}t}{\mathrm{e}}.\tag{7}$$

In Figure 7A, the best fit of experimental data calculated using Equation (7) is represented clearly as confirmed by the high value of the determination coefficient (*R*2) obtained for the linearized form of Equation (7) (Table 2, Figure 7B).

**Figure 7.** (**A**) Photodegradation of ibuprofen catalyzed by TiO2-coated blue glass; Ctcalc, values calculated using Equation (6); Ct exp, experimental values; error bars represent the standard deviations of three replicate experiments. (**B**) Trend of first-order linearized equation used for the calculation of kinetic parameters reported in Table 2.

In this case, the half-life can be calculated as

$$\mathbf{t}\_{1/2} = \text{L.r2/k.}\tag{8}$$

The value obtained was 575 min (Table 2), which is far from the half-life resulting from the experiment with TiO2 powder; nevertheless, it is satisfactory if we consider the high stability shown by IBP molecules not only in the darkness (for testing thermal and hydrolysis degradation), but also under light irradiation (photolysis degradation). Moreover, from Equation (8), it is possible to notice that, unlike the case of the second-order reaction, the half-life for the first-order reaction does not depend on the initial concentration of the reactant. This means that the reaction catalyzed by TiO2 immobilized on the active glass surface does not suffer from the same limitations encountered in the case of the second-order kinetics shown by IBP under the photoreaction catalyzed by TiO2 powder dispersion (light scattering, radical–radical recombination reactions, and dilution effect). Furthermore, it should be emphasized that (i) the number and persistence of derivatives was reduced in the case of coated active glass, and (ii) the time needed for an efficient degradation of the mother drug and its derivatives was approximately the same.

We have to mention that the degradation of IBP is also influenced by the pH value of the medium [30]. The production of hydroxyl radicals is generally increased in an alkaline medium, since high concentrations of OH− result in the formation of hydroxyl radicals, which are produced from the reaction of OH<sup>−</sup> with TiO2 on its surface's holes [31]. The pH value can also affect the charge on the catalyst particles; consequently, the electrostatic interactions between the charged surface of TiO2 and the pollutant molecules can be largely influenced, thus leading to a change in the adsorption level of these molecules on the catalyst surface and interfacial electron transfer [32].The most dominant factor affecting the adsorption of pollutant on catalyst surface is the catalyst zero-point charge (zpc), which is defined as the pH at which the surface of the catalyst has neutral charge [33].

For TiO2 P25 Evonik-Degussa, the zero-point charge value is 6.9. Therefore, the surface of TiO2 is positively charged in acidic media and negatively charged in basic media [33]. Accordingly, the effect of pH depends mainly on the type of the pollutant and the zero = point charge of the semiconductor. As IBP is weakly acidic in nature, it is expected to be negatively charged at pH higher than 3 [30], while the TiO2surface is positively charged at pH less than 6.9 [34]. Therefore, at pH = 4.5 where the photocatalytic experiment took place, the adsorption of IBP and, consequently, its photocatalytic oxidation were favored [30].

#### *2.4. Identification of Intermediate Photoproducts*

For the identification of IBP degradation byproducts, samples were collected at various time intervals and analyzed and identified by LC-FTICR MS system in the *m*/*z* range of 50–1000 in negative ionization mode. The results indicate the formation of two major photoproducts (Table 3). In addition, they reveal that the hydroxyl radicals attacked both the propionic acid and isobutyl substituent in IBP, resulting in the formation of two products, 2-[4-(1-hydroxyisobutyl) phenyl] propionic acid (2) and 4-(1-hydroxy isobutyl) acetophenone (3). Figure 8 depicts the proposed reaction pathways. The peak that appears at a nominal *m*/*z* value of 221, showing the formation of a mono-hydroxylated product of IBP, corresponds to 1-hydroxy IBP (2). Furthermore, product 2 was converted into another derivative with a nominal *m*/*z* value of 191, which corresponds to 4-(1-hydroxy isobutyl) acetophenone (3).


**Table 3.** Identification of ibuprofen and its photoproducts during photocatalytic degradation as deprotonated molecules, [M–H]−, by high-resolution LC-ESI-FTICR MS.

<sup>a</sup> Number used to identify each compound in the chromatograms of Figures 4 and 5. <sup>b</sup> Chromatographic retention time of compounds eluted under the experimental conditions described in Section 3. <sup>c</sup> Molecular formula of deprotonated compound. <sup>d</sup> Accurate *<sup>m</sup>*/*<sup>z</sup>* value of deprotonated molecules. <sup>e</sup> Mass error in parts per million <sup>=</sup> <sup>10</sup><sup>6</sup> × (accurate mass − exact mass)/exact mass.

**Figure 8.** By-products generated by TiO2 photocatalytic processes identified by LC-FTICR MS system in negative ion mode and proposed photodegradation pathway.

It is worth noting that, in the photodegradation method using TiO2 immobilized on the active glass surface, only one by-product, compound 2, was detected.

#### **3. Materials and Methods**

#### *3.1. Chemicals and Analytical Methods*

Ibuprofen (MW, 206.3 g·mol<sup>−</sup>1; pKa, 5.0) pure standard (purity, 99%) was purchased from Sigma Aldrich (Munich, Germany); acetonitrile, formic acid, and water for analysis were HPLC grade and purchased from Sigma Aldrich; TiO2 Degussa P-25 was a kind compliment from Evonik Industries (Steinheim, Germany); TiO2-coated active glass (Figure 9) was obtained as a gift from Pilkington (UK) (Sheel et al. 1998). PTFE (polytetrafluoroethylene) filters, 0.2 μm pore size, filter-Ø: 25mm, were purchased from Macherey-Nagel GmbH & Co. KG (Duren, Germany). Daily fresh working solutions were prepared using ultra-pure water from a bi-distilled purification system.

**Figure 9.** TiO2-coated blue glass.

To avoid microbial contamination, all glass apparatus was heat-sterilized by autoclaving for 60 min at 121 ◦C before use. Aseptic handling materials and laboratory facilities were used throughout the study to maintain sterility.

IBP concentrations were monitored using high-performance liquid chromatography (HPLC) (1200 series, Agilent Technologies, Santa Clara, USA) equipped with an Eclipse XDB-C18 (3 μm particle size, 4.6 × 150 mm) column (Phenomenex, Torrance, USA) using a diode array detector (DAD) at a wavelength of 230 nm. The mobile phase consisted of 40% of 1% formic acid solution/60% acetonitrile. The flow rate was 1.0 mL·min−1. Several aqueous solutions (from 0.5 to 25.0 mg·L−1) of IBP were filtered, and 20 μL of the filtrate was injected and analyzed. Peak areas vs. concentration of IBP were plotted, and the calibration curve was obtained with a determination coefficient (*R*2) of 0.9986. The limit of detection (LOD) of IBP for this method (using DAD) was 0.2 mg·L<sup>−</sup>1, and the limit of quantitation (LOQ) was 0.6 mg·L<sup>−</sup>1. The identification of IBP photoproducts was performed using the LC-FTICR MS system (Thermo Fisher Scientific, Bremen, Germany), in the same separation conditions. Negative ion ESI-MS mode was used for the detection of the compounds of interest. Full-scan experiments were performed in the ICR trapping cell in the range *m*/*z* 50–1000. Mass-to-charge ratio signals (*m*/*z*) were acquired as profile data at a resolution of 100,000 full width at half maximum (FWHM) at *m*/*z* 400. The limit of detection for mass spectrometric method was a few pmols.

The photodegradation experiments were performed using a solar simulator device Heraeus Sun-test CPS+ (Atlas, Chicago, USA), equipped with a 1500-W xenon arc lamp protected with a quartz filter (total passing wavelength: 280 nm < λ < 800 nm). The irradiation chamber was maintained at 20 ◦C by circulating water from a thermostatic bath and through a conditioned airflow.

#### *3.2. Characterization of the TiO2-Coated Active Glass*

The active glass was obtained via a coating process using a nanocrystalline film of TiO2 on a 4-mm-thin glass sheet. Some cross-sections obtained from theTiO2-coated active glass were analyzed. The scanning electron analysis for the TiO2-coated active glass was accomplished using a scanning electron microscope (SEM) of LEO model EVO50XVP, Carl Zeiss AG-EVO® 50 Series (Germany). The thin sections were grafted with 30-nm-thick carbon films. Semi-quantitative analyses of the

elemental composition of the different layers were obtained using a Ge ED Oxford-Link detector equipped with a super atmosphere thin window. Operating conditions of the SEM were as follows: 15 kV accelerating potential, 500 pA probe current, and about 10 mm working distance (WD). Thin sections of glass were prepared by the Department of Health and Environmental Science, Bari University. Samples were embedded in resin epoxy plugs and then polished.

#### *3.3. Photolysis Experiment*

Aqueous IBP solution of initial concentration 25 mg·L−<sup>1</sup> was prepared by dissolving a determined quantity of standard IBP in ultrapure water. The measured pH of the solution was 4.5. The photolysis treatment was carried out in a glass Pyrex®batch reactor closed at the top with a quartz cover. IBP solution (250 mL) was placed into the reactor; then, thereactor was placed into the irradiation oven inside the solar simulator, which reproduced the spectral distribution of natural solar irradiation. The IBP solution was continuously remixed during the experiment by magnetic stirring, and samples were taken (1 mL for each sample) at determined intervals and then analyzed using the HPLC system according to the analysis method in Section 3.1. Three experiments of direct photolysis were performed in triplicate.

#### *3.4. Photocatalysis Experiment with TiO2Powder*

A solution of ibuprofen was prepared as described in Section 3.3, but 50 mg of TiO2 powder (0.2 g·L−1) was added as the optimized amount in the reactor vessel. The aqueous suspension was mixed continuously in the dark for 2 h to ensure that the adsorption equilibrium of IBP on the catalyst surface was reached; then, the reactor was transferred into the solar simulator and exposed to solar irradiation, and samples were taken (1 mL for each sample) at determined intervals, then filtered and analyzed by HPLC. Three experiments were performed in triplicate.

#### *3.5. Photocatalysis Experiment with TiO2Immobilized on Active Glass*

Seven active glass sheets were placed vertically to cover the full surface of the inner wall of the reactor; then, the solution of IBP, prepared as in the previous two experiments, was added, and the reactor was transferred into the solar simulator and exposed to xenon lamp irradiation with continuous mixing; samples were taken (1 mL for each sample) at determined intervals, and then filtered and analyzed. Three experiments were performed in triplicate.

#### **4. Conclusions**

IBP is very stable under direct photolysis conditions due to its high chemical stability and low molar adsorption coefficient in the range of wavelengths provided by solar irradiation. On the other hand, our experiments showed that effective destruction of IBP and its photoproducts is possible by photocatalysis in the presence of TiO2 powder suspension or using TiO2immobilizedon the surface of active glass. Two intermediate photo-products were detected and identified by LC-FTICR MS. As there is substantial equivalence in the long-term efficacy of photocatalys is in the presence of TiO2 both as powder suspension and as glass coating, the use of active glass instead of TiO2 suspension could be a promising technique for the removal of pharmaceutical residues such as IBP and its photoproducts from aquatic environments not requiring the recovery of the catalyst after photodegradation. To increase the effectiveness of the technique described herein, a modification of TiO2immobilization on the glass surface using supports with a more complex geometry is essential.

**Author Contributions:** S.K. This work is a part of his PhD thesis (Chapter 4). Therefore all experiments and results and manuscript writing were referring to his own efforts; J.H.S. Supervision and reviewing results, reviewing the manuscript English writing; F.L. LCMS analysis; L.S. SEM analysis; S.A.B. Supervision and reviewing results, reviewing the manuscript English writing; R.K. Reviewing the manuscript English writing. All authors have read and agreed to the published version of the manuscript.

**Funding:** This work was supported by the European Union in the framework of the Project "Diffusion of nanotechnology-based devices for water treatment and recycling; NANOWAT" (ENPI CBC MED I-B/2.1/049, Grant No. 7/1997).

**Acknowledgments:** Many thanks to Jawad H. Shoqueir, the head of Soil and hydrology Lab at Al-Quds University, for his support to partially cover the publication fee from his own budget. Results reported in this article were partially presented by Samer Khalaf at the Second International Conference on Recycle and Reuse, 4–6 June 2014, Istanbul, Turkey and published in the book of abstracts.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


© 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

### *Article* **Photocatalytic Degradation of Chlorpyrifos with Mn-WO3**/**SnS2 Heterostructure**

#### **Charlie M. Kgoetlana, Soraya P. Malinga and Langelihle N. Dlamini \***

Department of Chemical Sciences, University of Johannesburg, Doornfontein Campus, P.O. Box 17011, Doornfontein, Johannesburg 2028, South Africa; kgoetlanacm@gmail.com (C.M.K.); smalinga@uj.ac.za (S.P.M.)

**\*** Correspondence: lndlamini@uj.ac.za; Tel.: +27-011-559-6945

Received: 21 May 2020; Accepted: 4 June 2020; Published: 21 June 2020

**Abstract:** Tungsten trioxide (WO3) is a photocatalyst that has gained interest amongst researchers because of its non-toxicity, narrow band gap and superior charge transport. Due to its fast charge recombination, modification is vital to counteract this limitation. In this paper, we report on the fabrication of Mn-doped WO3/SnS2 nanoparticles, which were synthesised with the aim of minimising the recombination rates of the photogenerated species. The nanomaterials were characterised using spectroscopic techniques (UV-Vis-diffuse reflectance spectroscopy (DRS), Raman, XRD, photoluminescence (PL) and electrochemical impedance spectroscopy (EIS)) together with microscopic techniques (FESEM-EDS and high resolution transmission electron microscopy selected area electron diffraction (HRTEM-SAED)) to confirm the successful formation of Mn-WO3/SnS2 nanoparticles. The Mn-doped WO3/SnS2 composite was a mixture of monoclinic and hexagonal phases, confirmed by XRD and Raman analysis. The Mn-WO3/SnS2 heterojunction showed enhanced optical properties compared to those of the un-doped WO3/SnS2 nanoparticles, which confirms the successful charge separation. The Brunauer–Emmett–Teller (BET) analysis indicated that the nanoparticles were mesoporous as they exhibited a Type IV isotherm. These nanomaterials appeared as a mixture of rectangular rods and sheet-like shapes with an increased surface area (77.14 m2/g) and pore volume (0.0641 cm3/g). The electrochemical measurements indicated a high current density (0.030 mA/cm2) and low charge transfer resistance (157.16 Ω) of the Mn-WO3/SnS2 heterojunction, which infers a high charge separation, also complemented by photoluminescence with low emission peak intensity. The Mott–Schottky (M-S) plot indicated a positive slope characteristic of an *n*–*n* heterojunction semiconductor, indicating that electrons are the major charge carriers. Thus, the efficiency of Mn-WO3/SnS2 heterojunction photocatalyst was monitored for the degradation of chlorpyrifos. The effects of pH (3–9), catalyst loading (0.1–2 g) and initial chlorpyrifos concentration (100 ppb–20 ppm) were studied. It was observed that the degradation was purely due to photocatalysis, as no loss of chlorpyrifos was observed within 30 min in the dark. Chlorpyrifos removal using Mn-WO3/SnS2 was performed at the optimum conditions of pH = 7, catalyst loading = 1 g and chlorpyrifos concentration = 1000 ppb in 90 min. The complete degradation of chlorpyrifos and its major degradation by-product 3,5,6-trichloropyridin-2-ol (TCP) was achieved. Kinetic studies deduced a second order reaction at 209 <sup>×</sup> <sup>10</sup>−<sup>3</sup> <sup>M</sup><sup>−</sup>1s−1.

**Keywords:** heterojunction; charge separation; photocatalysis; chlorpyrifos

#### **1. Introduction**

The fabrication and modification of photocatalysts has sparked interest amongst researchers due to their wide applications. Photocatalysts are used in applications ranging from water splitting, the degradation of pollutants in water, gas sensing and optoelectronic devices [1]. These can be *n*-type (electrons are the major charge carriers) or *p*-type (holes are the major charge carriers) photocatalysts [2]. The most widely studied photocatalysts are TiO2, WO3 (*n*-type) and ZnO, CdS (*p*-type) [3,4]. The photocatalytic efficiency of these materials is limited to a certain extent, primarily due to two major limitations. Firstly, they are prone to fast electron–hole recombination, which reduces the photocatalytic reactivity of the semiconductor. Secondly, they have wide band gaps that absorb only in the ultraviolet (UV) region, which accounts for 4% of the solar spectrum [5].

Modifications of photocatalysts to suit specific applications have been proposed. These include the use of metal dopants to form Schottky barriers and fusion with other semiconductor photocatalysts, resulting in heterojunctions [2]. The metal dopants that have been employed include magnesium (Mg), manganese (Mn), copper (Cu) and yttrium (Y) [6–9]. The metal doping of photocatalysts results in shifting the absorption band edge of the material to absorb the readily available visible region of the solar spectrum. They also separate photogenerated charges by forming electron traps, although a high concentration of the metal dopant may result in the creation of recombination centres, which leads to an increased recombination rate.

Heterojunctions that have shown enhanced optical and photocatalytic properties include BiVO4/WO3, CdS/ZnO and TiO2/SnO2 [10–13]. The formation of Type II heterojunctions using two different photocatalysts is sufficient to reduce the recombination rate of photogenerated charges. This occurs by the accumulation of photoexcited electrons in the conduction band (CB) of one semiconductor while photogenerated holes accumulate in the valence band (VB) of another semiconductor in the heterojunction system, which effectively leads to charge separation. Therefore, photo-oxidation and photo-reduction occur in different semiconductor surfaces of the heterojunction system due to the different migration points of the charges [14].

Tungsten trioxide (WO3) is a visible light photocatalyst with a band gap energy of 2.5–2.8 eV [15]. It is classified as an *n*-type semiconductor, wherein electrons are the major carriers. Due to its narrow band gap, WO3 absorbs light radiation in the visible range, and it has been used in a wide range of applications such as fuel production and combating water pollution [16,17]. This semiconductor exists in different polymorphs, which include monoclinic, triclinic, tetragonal and orthorhombic. The monoclinic phase of WO3 is the most stable and most photocatalytic compared to all the other phases.

Like most photocatalysts, WO3 suffers from limitations such as high electron–hole charge recombination. To overcome the intrinsic limitation of pristine WO3, different methods have been used, including metal doping and loading another semiconductor photocatalyst to form a heterojunction [10,18,19].

We, however, report the synthesis and characterisation of a material that fuses metal doping and a heterojunction (Mn-WO3/SnS2) that exhibits improved optical properties and minimises electron–hole recombination compared to current photocatalysts. Owing to the charge mobility in SnS2 to efficiently facilitate electron transfer to the WO3 CB, there results a high number of electrons for the oxidation reaction and separated holes that accumulate on the VB of SnS2. The Mn2<sup>+</sup> ions also separate charges by trapping electrons in the WO3, thereby increasing their lifetime, and act as reaction centres. This photocatalyst can be applied in water remediation, energy production and sensing. Thus, this study assessed the photo-efficiency of the heterostructure in the photodegradation of chlorpyrifos, an organophosphate pesticide.

Organophosphate pesticides have been used extensively in South Africa and the world at large due to their ability to combat a vast spectrum of pests [20,21]. Chlorpyrifos (O,O-diethyl O-[3,5,6, -trichloro-2-pyridyl] phosphorothionate) (CPF) is an organophosphorus pesticide extensively used in agricultural and domestic applications [22]. Chlorpyrifos agricultural application occurs throughout the year for a variety of fruits and vegetables [23]. It, however, does not readily dissolve in water, yet adsorbs strongly to soil particles.

Chlorpyrifos is an enzyme acetylcholinesterase inhibitor and persistent pesticide pollutant. It is a class II (moderately hazardous pesticide) pollutant, with a half-life of 60 days [24]. The pesticide is toxic to humans and other animals when ingested or inhaled; this is attributed to its lipophilic nature. It causes delayed peripheral neuropathy in humans and badly affect neuro-development in children at high doses [25,26].

Due to the numerous human and environmental effects caused by chlorpyrifos, different ways to remove this pesticide from the environment have been studied. These include advanced oxidation processes and biological treatment (with fungal and bacterial strains).

Bacterial strains have displayed high chlorpyrifos removal from water of up to 98% [27]. Though it is efficient, the method is strenuous, as bacteria require controlled specific conditions such as pH and temperature and a host for optimal function. On the other hand, Ismail et al. [24] in 2013 reported the use of advanced oxidation processes (AOPs) that yield 100% removal of chlorpyrifos by using 60Co γ-rays of 30–575 Gy [24].

However, γ-rays are harmful to human health; therefore, this led to the implementation of a better and safer method requiring the use of a photocatalyst to degrade chlorpyrifos under light irradiation. To date, zinc and titanium oxides have been used to degrade chlorpyrifos. The results were satisfactory, with up to 95% chlorpyrifos removal for TiO2 and 85% for ZnO under UV light [28]. However, the photocatalysts suffer from charge recombination and the use of UV light is not viable due to the insufficient amount of UV available (4%). Therefore, visible light-absorbing photocatalysts were discovered such as BiVO2, SnS2 and WO3.

To the best of our knowledge, no work has been reported to date on the fabrication of a metal-doped heterojunction (Mn-WO3/SnS2) photocatalyst.

#### **2. Results and Discussion**

#### *2.1. X-ray Di*ff*raction and Raman Analyses*

The phase and crystallographic properties of the nanomaterials were elucidated using XRD and Raman spectroscopy. Figure 1a shows the XRD pattern of WO3, which confirms the monoclinic nature of the WO3 (*m*-WO3). The *m*-WO3 was indexed and matched to the miller indices (002), (020), (200), (120), (112), (022), (202), (122), (222), (004), (040), (400), (042) and (420) (JCPDS Card No. 00-043-1035). Doping the *m*-WO3 with Mn2<sup>+</sup> did not distort the phase of the WO3, which implies that it had been intrinsically inserted into the WO3 crystal lattice as depicted in Figure 1b. The XRD pattern of the WO3/SnS2 heterojunction showed the presence of both monoclinic (*m*-WO3) and hexagonal (*h*-WO3) phases, and the hexagonal phase of SnS2 could be indexed (JCPDS Card No. 00-023-0677), as shown in Figure 1c. Again, the structural integrity of the manganese-doped WO3/SnS2 heterojunction (Figure 1d) was not distorted by the incorporation of Mn2<sup>+</sup> in the system. The average crystallite sizes of the nanomaterials were determined using the Debye–Scherrer equation and are tabulated in Table S1, all with an average size of 40 nm, with SnS2 having a crystallite size of less than 20 nm.

The nature of the phases was further confirmed with Raman analysis. Figure 2a illustrates Raman bands at 717 and 818 cm−<sup>1</sup> and less intense bands at 212 and 313 cm−<sup>1</sup> corresponding to O-W-O stretching and bending in the molecule, respectively, which confirms a monoclinic WO3; this finding was also reported by Simelane et al. 2017 and Xie et al. 2012 [3,15]. As in XRD analyses, the doping of Mn2<sup>+</sup> had no effect on the phase of WO3, as depicted in Figure 2b. No secondary bands resulting from Mn-oxides were observed. The heterojunction (WO3/SnS2) was successfully formed and confirmed by the Raman band at 317 cm−<sup>1</sup> corresponding to the A1g mode of hexagonal phase SnS2 as observed by Ma et al. 2015 (Figure 2c) [29]. Figure 2d displays the Raman band of Mn-WO3/SnS2 with no distortion due to Mn2<sup>+</sup> and SnS2. Therefore, the Raman band in Figure 2e corresponds to the pristine hexagonal phase of SnS2 due to the A1g band at 317 cm<sup>−</sup>1.

**Figure 1.** XRD patterns of (**a**) WO3, (**b**) Mn-WO3, (**c**) WO3/SnS2, (**d**) Mn-WO3/SnS2 and (**e**) SnS2.

**Figure 2.** Raman spectra of (**a**) WO3, (**b**) Mn-WO3, (**c**) WO3/SnS2, (**d**) Mn-WO3/SnS2 and (**e**) SnS2.

#### *2.2. Morphological Studies*

The morphological studies were conducted using microscopic techniques such as FESEM and HRTEM. Figure 3a is the FESEM image of pristine *m*-WO3 with rectangular sheets, rods and cubes, and the composition was confirmed by EDX (inset). The shapes of the nanomaterials did not change upon the insertion of Mn2<sup>+</sup> or the formation of the heterojunction as illustrated in Figure 3b,c and Figures S2 and S3. The EDX spectra displayed the elemental composition of the respective heterojunctions (WO3/SnS2) and Mn-WO3/SnS2 (inset).

**Figure 3.** FESEM images (inset is the corresponding energy-dispersive X-ray (EDX) spectrum) of (**a**) pristine WO3, (**b**) WO3/SnS2 and (**c**) Mn-WO3/SnS2, and TEM images (inset is the corresponding selected area electron diffraction (SAED) image) of (**d**) WO3, (**e**) WO3/SnS2 and (**f**) Mn-WO3/SnS2.

The HRTEM image of *m*-WO3 also showed rectangular sheets and rods (Figure 3d) and was further elucidated using SAED (inset) obtained through a 1–10 zone axis. The spots were indexed to (002), (220) and (112) corresponding to monoclinic WO3 as confirmed by XRD analysis. The HRTEM images (Figure 3e,f) displayed rectangular rods and sheet-like shapes as observed in Figure 3d, which implies that no shape distortion had occurred through metal doping and the formation of the heterojunction. The SAED image (inset) displays spots and rings characteristic of monoclinic WO3 and the SnS2 hexagonal phase, respectively. Furthermore, the SAED image (inset) illustrates spot (202, 200) and ring (101, −103) indices corresponding to the WO3 monoclinic phase and SnS2 hexagonal phase, respectively, captured through a 0–10 zone axis using the CrysTBox software [30] (Figure 3f). All the SAED indices correspond to the reported XRD patterns, which further confirms the successful formation of our nanomaterial. In the Mn-WO3/SnS2, the estimated percentage of Mn was 2.5%, with 47.5% of WO3 and 50% of SnS2.

#### *2.3. Optical Properties*

Ultraviolet-visible spectroscopy in diffuse reflectance mode was used to determine the optical properties of the synthesised nanoparticles. All the synthesised nanomaterials showed a shift of absorption to be in the visible region, which is in abundance. Pristine *m*-WO3 displayed a band gap of 2.71 eV, with a corresponding absorption wavelength of 466 nm, as shown in Figure 4a and Figure S4, respectively. The value obtained agrees with the value reported by Simelane et al. in 2017 [15]. The insertion of Mn2<sup>+</sup> in the *m*-WO3 lattice introduced impurities and thus a shift in the Fermi level below the conduction band, promoting the red-shifting of WO3 on the absorption spectrum (Figure 4b);

this was observed by Harshulkhan et al. in 2017 using magnesium as a dopant [6]. The band gap of the WO3 was reduced upon the insertion of Mn and decreased further after the formation of a heterojunction with SnS2 (Figure 4a,b,d,e). The Mn-doped heterojunction had the lowest band gaps (2.08 eV and 2.34 eV) amongst the nanomaterials (the others were WO3, SnS2, Mn-WO3 and WO3/SnS2), which correspond to a high light absorption wavelength (red-shift) (Figure S4); this was due to visible light absorption enhancement by both the Mn ion and SnS2.

**Figure 4.** Tauc plots indicating the band gaps of (**a**) WO3, (**b**) Mn-WO3, (**c**) SnS2, (**d**) WO3/SnS2 and (**e**) Mn-WO3/SnS2.

The diagram in Figure 5 illustrates the change in the band edge potential of the synthesised semiconductor photocatalysts. The valence band edge potential (*E*VB) and the conduction band edge potential (*E*CB) were calculated using Equations (5) and (6). A slight decrease in both *E*CB and *E*VB was observed during the introduction of Mn and fusion with SnS2 (Mn-WO3/SnS2). The conduction band edge potential shifted to be more positive (by 0.2 eV), and the valence band edge potential moved to a less positive potential (by 0.2 eV); this was due to the insertion of an ion with a high ionic radius, which reduces the band gap by pulling the band edges closer, resulting in band edge shifts. The change in the position of the band edges enhances the absorption wavelength of the material. The heterojunction (Mn-WO3/SnS2) enhances charge separation by the movement of electrons from the SnS2 CB to the WO3 CB through the interface, thereby leaving holes in the VB of the SnS2. This effectively separates the electrons and holes as they accumulate in the CB of WO3 and the VB of SnS2, respectively.

**Figure 5.** Diagram of the band *gap*, *valence band* and *conduction* band *edge positions* vs. the NHE of the (**a**) WO3, (**b**) SnS2, (**c**) Mn-WO3, (**d**) WO3/SnS2 and (**e**) Mn-WO3/SnS2 photocatalysts.

#### *2.4. Electrochemical and Photoluminescence Measurements*

The electrochemical impedance spectroscopy (EIS) measurements were carried out to study the interfacial reactions occurring between the photoelectrode and the electrolyte. Figure 6a illustrates the EIS spectrum (Nyquist plot) with a suppressed semicircle with a large diameter. At low frequency, the current density is in phase with the potential deviation of the system, resulting in a straight line at an angle of 45◦ to the *X*-axis. The large diameter of the semicircle at high frequency corresponds to the high charge transfer impedance of WO3. This relates to the high charge recombination rate as observed in Figure 6a. The charge transfer impedance was reduced after WO3 was doped with the Mn2<sup>+</sup> ion (Figure 6b), due to the reduced charge recombination rate and increased charge mobility. This was due to the Mn2<sup>+</sup> ions acting as charge collection sites, thereby serving as an electrical conduction pathway, allowing ion/electron mobility on the electrode [7]. The small diameter of the semicircle of the Mn-WO3/SnS2 spectrum indicates decreased electrode–electrolyte charge-transfer resistance/impedance compared to that in the WO3/SnS2, Mn-WO3, SnS2 and WO3 in the 0.1 M Na2SO4 electrolyte. The sloping straight line in the low-frequency region corresponds to oxygen diffusion within the electrode (Figure 6a). The low charge transfer resistance of Mn-WO3/SnS2 arises from the enhanced charge carrier separation induced by the Mn2<sup>+</sup> ion dopant and SnS2 semiconductor heterojunction with WO3. The charge transfer resistance and recombination rate decreased for WO3, Mn-WO3, SnS2 and WO3/SnS2, with the lowest rate observed in the Mn-WO3/SnS2 (Figure 6a,b).

**Figure 6.** (**a**) Electrochemical impedance spectra (Nyquist plot), (**b**) photoluminescence spectra, (**c**) linear sweep voltammetry, and (**d**) Mott–Schottky plot of (a) WO3, (b) Mn-WO3, (c) SnS2, (d) WO3/SnS2 and (e) Mn-WO3/SnS2.

Photoluminescence (PL) measurements (Figure 6b) support the EIS findings that the WO3 has a strong PL intensity, which indicates high charge carrier recombination, and it was reduced by the introduction of Mn and fusion with SnS2. A decrease in PL intensity was observed in the Mn-WO3/SnS2, indicating low charge carrier recombination, which implied that it would be a good photocatalyst in photocatalytic applications. This is attributed to the longer charge carrier lifetimes and enhanced charge carrier mobility provided by the Sn-S bond, thereby minimising the electron–hole recombination.

Upon the introduction of Mn2<sup>+</sup> and SnS2, the photocurrent density of WO3 was observed to be improved by up to 0.030 mA/cm<sup>2</sup> for Mn-WO3/SnS2 NPs (Figure 6c). This implied that there was high electron flow between the photocatalyst and the electrolyte produced from the photocatalyst upon light irradiation. The Mn2<sup>+</sup> acts as an electron sink and reaction side, which in turn supplies electrons for interfacial reactions, and upon illumination, SnS2 helps in the production of electrons and their separation from holes, which increases the current density.

Mott–Schottky plots were used to study the interfacial capacitance of the nanomaterials. The positive slopes obtained from Figure 6d confirmed that the synthesised nanomaterials are all *n*-type semiconductors, which use electrons as major charge carriers. The positive slope for the heterojunction WO3/SnS2 and Mn-WO3/SnS2 NPs further inferred the formation of an *n–n* type heterojunction system. Upon the introduction of the Mn2<sup>+</sup> and formation of the heterojunction, there was no significant change in the slope of the curves.

The flat-band potential (*V*fb) was obtained by extrapolating a line on the slope of the graph to the *x*-intercept (1/C2=0). The flat-band potentials were found to be 0.214 V, 0.159 V, <sup>−</sup>0.209 V, <sup>−</sup>0.103 V and −0.039 V, corresponding to WO3, Mn-WO3, SnS2, WO3/SnS2 and Mn-WO3/SnS2, respectively. The flat-band potential in *n*-type semiconductors corresponds to the bottom of the conduction band of the semiconductor photocatalyst, which was observed to decrease upon doping and the formation of the heterojunction (Figure 6d). The obtained flat-band potential (*V*fb) values were found to correspond to the calculated conduction band edge potentials (*E*CB) from UV-Vis DRS.

The Randles equivalent circuit model was used to fit the obtained EIS data. The Randles equivalent circuit models (Figure 7a,b) corresponding to the graphs show that the impedance was a contribution of three forms of resistance, namely the solution resistance, the electrode resistance due to the film composition of the nanomaterials, and charge-transfer resistance occurring at the electrolyte–electrode interface.

**Figure 7.** Randles equivalent circuit models corresponding to (**a**) WO3 and Mn-WO3, and (**b**) SnS2, WO3/SnS2 and Mn-WO3/SnS2.

The Warburg impedance is due to solid-state ion diffusion during the electrochemical reaction. The Warburg element manifests itself in EIS spectra as a straight line with a slope of 45◦ in the low-frequency region. Different slopes of the straight-line part in the low-frequency region indicate that the electrodes have different Warburg impedances and solid-state ion diffusion behaviors.

The equivalent circuit model was obtained after fitting the data to the Randles model: R1 is the solution resistance, R2 is the thin layer resistance, R3 is the charge transfer resistance, W2 is the Warburg resistance and Q1, Q2 and Q3 are the constant phase elements. The Warburg impedance relates to solid-state ion diffusion during the electrochemical reaction in the solution. This favours photocatalytic activity by utilising the separated charges during the reaction and, consequently, reduces charge recombination. The slope of the Warburg transition line also indicates the reactivity of the nanoparticles. Furthermore, the charge transfer impedance was found to be 631.80, 498.50, 310.55, 173.65 and 157.16 Ω for WO3, Mn-WO3, SnS2, WO3/SnS2 and Mn-WO3/SnS2, respectively.

#### *2.5. BET Analysis*

The BET analysis revealed that the nitrogen adsorption isotherms obtained for the nanoparticles were Type IV isotherms, according to the IUPAC (International Union of Pure and Applied Chemistry) classification (indicated in Figure S6). A Type IV isotherm is typical of mesoporous materials (IUPAC definition: pore size 2–50 nm), suggesting that the nanomaterials consist of agglomerates. The Mn-WO3/SnS2 nanoparticles were found to have the highest BET surface area (77.14 m2/g) and pore volume, of 0.0641 cm3/g, compared to pristine materials (Table 1). This suggests that Mn-WO3/SnS2 would have improved photocatalytic activity due to the adsorption capacity provided by its large specific surface area during photocatalysis. The large pore volume would allow the efficient trapping of pollutants during adsorption for degradation to take place.


**Table 1.** The specific surface area and pore volume of NPs.

#### *2.6. Surface Charge of Nanoparticles*

The stability of the nanomaterials in suspensions was studied using the electrophoretic light scattering technique. The zeta potentials of the nanomaterials are illustrated in Figure 8. showing a steady but gradual change in zeta potential from positive to negative as the pH increased from 2 to 11 for all the photocatalysts (Figure 8).

**Figure 8.** Surface charge of the nanoparticles.

The point of zero charge (pzc) for pristine WO3 was observed at pHpzc 2.5, which corresponds to what is reported in the literature.

A slight shift of the pzc to higher pH was observed for Mn-WO3 (pHpzc = 3.2). The shift is due to the substitution of W6<sup>+</sup> by Mn2<sup>+</sup> metal ions, consequently changing the overall charge of the material. Therefore, species adsorbed onto the surface of the photocatalyst change the surface charge and shift the point of zero charge of the suspended nanoparticles.

The point of zero charge for pristine SnS2 was found to be at pH 5.5, as reported in literature. Furthermore, the heterojunction (WO3/SnS2) displayed a point of zero charge (2.7) at a lower pH than SnS2 but higher than WO3; this was attributed to synergistic effects from both counterparts (WO3 and SnS2) in the heterojunction.

Furthermore, introduction of Mn in the heterojunction (Mn-WO3/SnS2) shifted the point of zero charge to 2.1, much lower than for all the other photocatalysts.

#### *2.7. Degradation of Chlorpyrifos*

The photodegradation of chlorpyrifos using the synthesized nanoparticles is showed in Figure 9. The degradation profile for chlorpyrifos indicated an increase in removal by the nanoparticles from the WO3 to Mn-WO3/SnS2 photocatalysts. The Mn-WO3/SnS2 nanoparticles showed high removal of chlorpyrifos due to high charge separation and lower charge impedance. Therefore, Mn-WO3/SnS2 represented the best performing photocatalyst with up to 95.90% chlorpyrifos removal, calculated using Equation (10).

**Figure 9.** Degradation of chlorpyrifos (1000 ppb) using different photocatalysts at pH = 5 and 0.1 g of photocatalyst.

Figure 10 displays the percentage removal of chlorpyrifos in water within a period of 60 min. The removal efficiency for chlorpyrifos by using the nanoparticles resulted in 56.80%, 60.20%, 75.00%, 84.88% and 95.90% removal for WO3, Mn-WO3, SnS2, WO3/SnS2 and Mn-WO3/SnS2, respectively.

**Figure 10.** Percentage removal for chlorpyrifos (1000 ppb) using 0.1 g of (**A**) WO3, (**B**) Mn-WO3, (**C**) SnS2, (**D**) WO3/SnS2 and (**E**) Mn-WO3/SnS2.

The reaction kinetics correspond to the percentage chlorpyrifos removal. The rate constants (K) of the reactions using the respective photocatalysts are presented (Figure 11), which were 9.3 <sup>×</sup> <sup>10</sup>−<sup>3</sup> <sup>M</sup><sup>−</sup>1min−<sup>1</sup> and 209 <sup>×</sup> <sup>10</sup>−<sup>3</sup> <sup>M</sup><sup>−</sup>1min−<sup>1</sup> for WO3 and Mn-WO3/SnS2, respectively.

**Figure 11.** Rate constants of (**A**) WO3, (**B**) Mn-WO3, (**C**) WO3/SnS2, (**D**) Mn-WO3/SnS2 and (**E**) SnS2.

The photodegradation reaction was fitted to Equation (10) from which the rate constant k(M−1min−1) was calculated from the gradient of the plot of 1/[C] against time (t). The reaction kinetics leading to the determination of the rate constant followed a second order reaction pathway.

The rate constants were 9.3 <sup>×</sup> 10−<sup>3</sup> M−1min−1, 14.3 <sup>×</sup> 10−<sup>3</sup> M−1min−1, 25.0 <sup>×</sup> 10−<sup>3</sup> M−1min−1, 47.4 <sup>×</sup> <sup>10</sup>−<sup>3</sup> <sup>M</sup><sup>−</sup>1min−<sup>1</sup> and 209.5 <sup>×</sup> <sup>10</sup>−<sup>3</sup> <sup>M</sup><sup>−</sup>1min−1, corresponding to WO3, Mn-WO3, SnS2, WO3/SnS2 and Mn-WO3/SnS2, respectively (Figure 11).

The linear plot for the Mn-WO3/SnS2 nanoparticles kinetic studies is illustrated in Figure 12. The rate constant is 209.5 <sup>×</sup> <sup>10</sup>−<sup>3</sup> <sup>M</sup><sup>−</sup>1min−1, and the R2 is 0.9656.

**Figure 12.** Photodegradation kinetics for chlorpyrifos using Mn-WO3/SnS2.

#### *2.8. E*ff*ect of pH on the Photocatalytic Activity*

The surface charge of the nanoparticles in a suspension is influenced by the pH of the solution. The photodegradation of chlorpyrifos using Mn-WO3/SnS2 nanoparticles increased with an increase in the pH of the initial solution, as illustrated in Figure 13. The point of zero charge for Mn-WO3/SnS2 is at pHpzc = 2.13 and above that is increasingly negative, as displayed by Equations (1) and (2).

$$\rm{M-OH} + \rm{H}^+ = \rm{M-OH}\_2^+ \ (\rm{pH} < \rm{pzc}) \tag{1}$$

$$\text{M-OH} = \text{M-O}^{-} + \text{H}^{+} \text{ (pH} > \text{pzc)} \tag{2}$$

The increase in the removal was also caused by the increase in the level of deprotonation of the nanoparticles at high pH, which influences the negative charge on the surface of the photocatalyst, consequently leading to high chlorpyrifos adsorption. That was also favoured by the positive charge of chlorpyrifos in alkaline solutions from pH 5, as reported in literature. There is a transfer of holes from the inner part of the nanoparticles to the surface, whereby OH− ions scavenge photogenerated holes and therefore yield very oxidative species such as •OH radicals. The percentage removal of chlorpyrifos achieved in 60 min was 85.6%, 94.3%, 99.8% and 99.0% at pH 3, pH 5.8, pH 7 and pH 9, respectively (Figure 13). Therefore, pH 7 was the optimum pH for chlorpyrifos removal using Mn-WO3/SnS2 nanoparticles and was used in the next sections. Hou et al. [31] in 2018 also reported pH 7 for optimum chlorpyrifos removal [20].

The increase in the removal was due to the increased electrostatic attraction between the photocatalyst and the chlorpyrifos that occurs when the pH is increased [28]. This causes an easy surface attachment, which implies that holes can oxidize chlorpyrifos directly and creates hydroxyl and superoxide radicals for further oxidation. As pH increased, the surface charge of the nanoparticles

also became more negative, which caused increased electrostatic attraction between the nanoparticles and chlorpyrifos.

**Figure 13.** Degradation of 1000 ppb chlorpyrifos using 0.1 g of Mn-WO3/SnS2 at different pH values.

#### *2.9. E*ff*ect of Initial Concentration*

The effect of initial chlorpyrifos concentration on the photocatalytic removal was studied, and the results are shown in Figure 14. The highest removal of 99.99% was achieved at a 100 ppb chlorpyrifos concentration, followed by 99.95% at 1000 ppb, compared to 94.40%, 87.51% and 84.38% at 5 ppm, 10 ppm and 20 ppm, respectively. The concentration of 1000 ppb was chosen as the best, because it is the highest concentration for which a high percentage removal was achieved.

**Figure 14.** Effect of the initial concentration on the removal of chlorpyrifos (1000 ppb) at pH 7 using 0.1 g of Mn-WO3/SnS2.

The decrease in the removal of chlorpyrifos was alluded to the opacity caused by the high chlorpyrifos concentration, which prevented the photocatalyst from utilising the irradiated light to produce reactive species for degradation. Again, the high concentration scatters the light, thereby inducing screening effects [28].

#### *2.10. E*ff*ect of Initial Photocatalyst Loading*

The initial photocatalyst loading's effect on the photoactivity was studied, and the results are presented in Figure 15. The photoactivity of Mn-WO3/SnS2 increased when 0.5 g of photocatalyst was used, then further increased when the photocatalyst loading was 1 g. The increase is a result of increased reactive surfaces, which further increase the rate and amount of chlorpyrifos removal [28].

**Figure 15.** Effect of the initial photocatalyst loading on the photodegradation of chlorpyrifos.

A high concentration of nanoparticles results in agglomeration, which further causes light scattering and screening effects, which reduce the specific activity of the photocatalyst. This further causes opacity, which prevents the further illumination of the photocatalyst.

Therefore, a decrease in chlorpyrifos removal was observed when 2 g of Mn-WO3/SnS2 was used, reaching up to 85%. This is compared to 0.1 g, 0.5 g and 1 g removing up to 99.95%, 99.98% and 99.99%, respectively. Thus, 1 g was the best performing, as it reached 98% removal within 30 min of reaction time.

#### *2.11. Mechanistic Pathway*

The mechanistic and proposed degradation pathway was evaluated, and the results are shown in Figure 16. The products obtained were 3,5,6-trichloropyridin-2-ol (TCP) and O,O-dihydrogen phosphorothioite. Only the O,O-dihydrogen phosphorothioite compound and no other by-product was observed, which implies that there was a complete degradation of chlorpyrifos and TCP in the synthetic water (Figure S7).

**Figure 16.** Proposed degradation pathway for chlorpyrifos.

#### **3. Materials and Methods**

#### *3.1. Materials*

Tungsten boride (WB) (≥97.0%), Tin(IV) chloride pentahydrate (SnCl4.5H2O, (98%)), manganese(II) chloride tetrahydrate (MnCl2•4H2O, (≥98%)), nitric acid (≥65%, Puriss), poly (vinylidene fluoride) (PVDF), *N*-methyl-2-pyrrolidone (NMP), silver paste, sodium sulphate (Na2SO4), silver/silver chloride (Ag/AgCl) electrode, sodium sulphide (Na2S), chlorpyrifos (PESTNATAL, 99.9%), methanol and formic acid were all supplied by Sigma-Aldrich (Pty) Ltd., Johannesburg, South Africa. The chemicals were used as received.

#### Synthesis of Nanomaterial

Pristine WO3 NPs were synthesised following the method developed by Xie et al. (2012) with slight modification [3]. Tungsten boride (4.12 mmol) was dissolved in 1 M HNO3 (56.0 mL) under constant stirring and then transferred into a 100.0 mL Teflon-lined stainless steel autoclave. The autoclave was then sealed and placed in an oven at 190 ◦C for 12 h with a heating rate of 16 ◦C per hour; thereafter, a yellow solution was obtained. The yellow solution was further centrifuged and washed with deionised water and dried at 100 ◦C for 12 h in an oven, resulting in a yellow solid product of WO3. The same procedure was followed to obtain Mn-WO3 via the one-pot synthesis of MnCl2•4H2O (10.00 mmol) with WB (4.12 mmol) in 56.0 mL of HNO3.

Pristine SnS2 NPs were synthesised by dissolving 2.73 mmol of SnCl4.5H2O in 40.0 mL of deionised water under continuous stirring at 60 ◦C for 10 min. Thereafter, Na2S (2.73 mmol) was added to the solution, and the mixture was then stirred for 10 min. The final mixture was then transferred into a 100.0 mL Teflon-lined stainless steel autoclave and heated at 180 ◦C for 12 h with a heating rate of 15 ◦C/h. The resultant solution was then centrifuged and washed with deionised water and dried at 100 ◦C for 12 h to obtain SnS2 nanoparticles.

The heterojunction WO3/SnS2 was synthesised stepwise using a hydrothermal method. The first step was adapted from the method for synthesising pristine WO3 and followed by the synthesis of SnS2 NPs on the surface of the dispersed WO3 NPs in 40.0 mL of deionised water. The SnS2 synthesis was adopted from the synthesis of pristine SnS2 NPs to obtain the heterojunction (WO3/SnS2).

Furthermore, the synthesis of the Mn-doped heterojunction WO3/SnS2 was carried out using hydrothermal treatment in a multistep method. Firstly, MnCl2•4H2O (10.0 mmol) and tungsten boride (4.12 mmol) were dissolved in 1 M HNO3 (56.0 mL) under constant stirring, and thereafter, the same procedure as in the synthesis of WO3 was followed to obtain a yellow solid product of Mn-WO3. Furthermore, Mn-WO3 (1.74 mmol) was dispersed in 40.0 mL of deionised water under continuous stirring and heating at 60 ◦C using a magnetic stirring hotplate. This was followed by the synthesis of SnS2 NPs on the surface of Mn-WO3 NPs adopted from the synthesis of pristine SnS2 nanomaterials to form Mn-doped WO3/SnS2 heterojunction composite nanoparticles.

#### *3.2. Characterization Techniques*

The synthesised nanoparticle phases were characterized using X-ray powder diffraction (XRD) (*PANalytical X'Pert* Pro-MPD Powder Diffractometer, Almelo, Netherlands) with CuKα radiation (0.1540 nm) and a monochromator beam in a 2θ scan range from 20◦–80◦. The instrument power settings used were 40 kV and 40 mA with a step size of 2θ (0.0080) and a scan step time of 87.63 s. The average crystallite size was calculated using the Debye–Scherrer equation, Equation (3):

$$L = \frac{K\lambda}{\beta \cos \theta} \tag{3}$$

where β is the full width at half maximum, λ is the X-ray wavelength (0.1541 nm) for CuKα, *K* = 0.89, and θ is the diffraction angle.

Raman spectroscopy (RamanMicro™ 200 PerkinElmer Inc., Waltham, MA, USA) with a single monochromator, a holographic notch filter and a cooled TCD, was used to detect and characterise the polymorphic forms of the NPs. The Raman spectra of the NPs were measured in a back-scattering geometry using an Ar-ion laser line (514.5 nm). Dark-field imaging was used with a power output of below 0.5 mW and an exposure time of 4.0 s. The morphological properties of the NPs were examined using high resolution transmission electron microscopy (HRTEM) (JOEL-TEM 2010) at an acceleration voltage of 200 kV. The ethanol-dispersed nanoparticles were deposited on a carbon-coated copper grid. Furthermore, selected area electron diffraction (SAED) images of the nanoparticles were captured and indexed using the CrysTBox software [29]. A field emission scanning electron microscope (FESEM) (TESCAN Vega TC instrument with VEGA 3 TESCAN software; TESCAN, Brno, Czech Republic) coupled with energy-dispersive X-ray (EDX) operated at 5.0 kV under a nitrogen gas atmosphere was used to further study the morphology and the elemental composition of the NPs. The optical properties were investigated using a UV-Vis spectrophotometer (Shimadzu UV-2450, Shimadzu Corporation, Kyoto, Japan) using diffuse reflectance spectroscopy (DRS) and BaSO4 as the reference material. The band gap (*Eg*) of the nanomaterials and a graph of (α*h*ν) against photon energy (*h*ν) was extrapolated following Equation (4):

$$
\alpha \text{lrb} = A(\text{lrb} - E\_{\%})^{\text{n/2}} \tag{4}
$$

where α is the absorption coefficient, *h*ν is the energy of the incident photon, *A* is a constant, and *Eg* is the band gap energy.

The value of *n* depends on the semiconductor transition type, which is a direct transition when *n* equals 0.5 and an indirect transition when *n* equals 2. The valence band edge potential (*EVB*) and the conduction band edge potential (*ECB*) were calculated using Equations (5) and (6):

$$E\_{CB} = \chi - E^{\text{e}} - 0.5E\_{\%} \tag{5}$$

$$E\_{VB} = E\_{CB} + E\_{\mathcal{X}} \tag{6}$$

where *ECB* and *EVB* are the conduction and valance band edge potentials, respectively; χ is the electronegativity of the semiconductor (the geometric mean of the electronegativities of all the constituent atoms); *E*<sup>e</sup> is the energy of free electrons on the hydrogen scale (4.5 eV); and *Eg* is the band gap energy of the semiconductor.

The photoluminescence spectra of the nanomaterials were obtained using a PerkinElmer fluorescence spectrometer (Model LS 45, PerkinElmer Inc., Waltham, MA, USA). A 300 W xenon lamp was used as a light source. The spectra were obtained at an excitation wavelength of 319 nm. The excitation and emission wavelengths were set at 319 nm and 605 nm, respectively. Specific surface area and pore volumes were determined using the Brunauer–Emmett–Teller (BET) method. Nitrogen was used as the adsorbate, and the nitrogen adsorption isotherms of the samples were obtained at 77K using a Micromeritics ASAP 2020 adsorption analyser (Micromeritics Instrument Corporation, Norcross, Georgia, USA). The samples were degassed before the analysis at 100 ◦C for 10 h. The pore volume was calculated from the amount of nitrogen adsorbed at the relative pressure (*P*/*P*o) of 0.980.

#### *3.3. Electrochemical Measurements*

The electrochemical measurements were conducted using a potentiostat (Gamry Interface 1000 potentiostat, Gamry Instruments, Philadelphia, PA, USA) in a standard three-electrode system employing Ag/AgCl (3.0 M KCl) as the reference electrode and Pt wire as the counter electrode. The working electrodes were the prepared nanomaterial mixed with polyvinylidene fluoride (PVDF) as a binder in a 10:1 ratio respectively, dispersed in 1 mL of N-methylpyridinium (NMP) solution and ultrasonicated for 30 min to obtain a homogeneous mixture. The obtained homogeneous mixture was drop casted onto the fluorine-doped titanium oxide (FTO-glass) substrate forming a thin film. The prepared electrodes were heated at 80 ◦C for 12 h in air. A copper wire was thereafter attached using a silver paste for charge transfer to the potentiostat from the paste and dried in air for 24 h.

The prepared electrodes were then applied in a three-electrode system for electrochemical impedance spectroscopy (EIS) at a frequency range of 10 kHz to 0.1 Hz at an AC voltage of 10 mV rms and DC voltage of 0.45 V vs. Ag/AgCl. The current density of the working electrode was determined by running a linear sweep voltammetry scan at a scan rate of 50 mV/s. The flat-band potential (*V*fb) values of the nanomaterials were obtained from Mott–Schottky plots (Equation (7)) at a frequency of 1000 Hz under the applied voltages of −2 to 2 V and a step voltage of 0.1 V.

$$\frac{1}{C^2} = \frac{2}{\left(\varepsilon \varepsilon\_o A^2 \varepsilon N\_D\right)} \left[ V - V\_{fb} - \left(\frac{k\_b T}{\varepsilon}\right) \right] \tag{7}$$

where *C* is the interfacial capacitance, *A* is the surface area of the electrode, *N*<sup>D</sup> is the donor density, *V* is the applied potential, and *Vfb* represents the flat-band potential. The temperature with dielectric constant and permittivity of free space are represented as *T*, ε and ε0, respectively. The charge of the electron (*e*) is 1.602 <sup>×</sup> 10−<sup>19</sup> C, and the Boltzmann constant (*k*B) is 8.617 <sup>×</sup> 10−<sup>5</sup> eV·K−1. All the electrochemical measurements were conducted in 0.1 M sodium sulphate (Na2SO4) solution as the electrolyte, and the values for the electrode potentials were recorded with reference to Ag/AgCl. A 300 W xenon lamp was used as the light source.

#### *3.4. Surface Charge*

Surface charge measurements were obtained using electrophoretic light scattering (ELS) with a Zetasizer NanoZS (Malvern) instrument. Zeta potential measurements were obtained using electrophoretic light scattering (ELS) to understand the surface charge of the nanomaterials as a function of the pH of the solution. The nanomaterials were suspended at 30 mg/L in deionized (DI) water. The pH of the suspensions was adjusted to a pH range of 2–10 using 1M NaOH and 1M HCl.

#### *3.5. Degradation of Chlorpyrifos*

#### 3.5.1. Chlorpyrifos Standard Preparations

A stock solution of Chlorpyrifos (0.01 g) was prepared in 1 L of deionized water, followed by a serial dilution to make 75, 50, 25, 12.5, 6.25, and 3.125 ppb solutions. The prepared solutions were thereafter transferred into 2 mL LC-MS vials, and 1 mL of deionized water was added. The working solution was maintained at pH 5.

#### 3.5.2. Photocatalytic Degradation of Chlorpyrifos

The photocatalytic activity of the nanomaterials was tested through the photodegradation of chlorpyrifos in synthetic water samples under visible light irradiation (Photoreactor, Lelesil Innovative Systems). The volume of the working solution was kept at 500 mL of chlorpyrifos solution. Initially, the concentration of the chlorpyrifos solution was 1 ppm and a photocatalyst loading of 0.1 g was used at pH 5. The photodegradation reaction occurred under continued magnetic stirring for 90 min under regulated temperatures of 20–25 ◦C, subjected to a cooling jacket using ice cubes.

The photocatalyst suspension containing chlorpyrifos was kept in the dark for 30 min before irradiation to allow equilibration. The samples were collected from the batch reaction before and after irradiation at set time intervals (10 min, 10 mL aliquots), filtered through a 0.45 μm PTFE membrane filter and transferred into a 2 mL LC-MS sample vial for analysis.

Furthermore, the optimization of reaction conditions such as the pH, initial chlorpyrifos concentration and initial photocatalyst loading were carried out. Therefore, the pH of the chlorpyrifos solution was adjusted to 3, 5, 7 and 9; the initial concentration of the pesticide (chlorpyrifos), to 100 ppb, 1 ppm, 5 ppm, 10 ppm and 20 ppm; and the photocatalyst loading, to 0.1, 0.5, 1 and 2 g.

#### *3.6. LC-MS Measurement*

Samples were analyzed using a triple quad UHPLC-MS/MS 8030 (Shimadzu Corporation) to monitor the removal of chlorpyrifos. The LC-MS/MS was fitted with a Nexera UHPLC upgrade with the capability to obtain 500 multiple reaction monitoring readings per second. The oven was equipped with a RaptorTM ARC-18 column (Restek Corporation) with a 2.7 μm pore diameter and length of 100 mm × 2.1 mm, maintained at 40 ◦C. The mobile phase consisted of 0.1% formic acid in water/methanol (9:1%, v/v) at a flow rate of 0.200 mL/min with a 10 μL injection volume. The ion source was electrospray ionisation (ESI) and was operated in positive mode. Meanwhile, LC-MS/MS data for the degradation intermediates were obtained after the full scan mode was run for 12 min at flow rate of 0.3 mL/min.

The percentage removal of chlorpyrifos from the synthetic water samples was calculated using Equation (8) below:

$$\% \text{ chlorpyryfos removal} = \left(1 + \frac{\text{C}}{\text{C}\_0}\right) \times 100 \tag{8}$$

where *C*<sup>0</sup> is the initial concentration and *C* is the final concentration of chlorpyrifos. The degradation products were determined by analysing the samples for a period of 60 min. The degradation pathway was then deduced from the mass/ion ratio obtained from the MS spectrum. The reaction kinetics of chlorpyrifos degradation were studied, the results were fitted to a second order model fitted, and a plot based on the calculated (1/[*C*]) versus reaction time was obtained following Equation (7).

$$\frac{1}{[\mathbb{C}]\_t} = kt + \frac{1}{[\mathbb{C}]\_0} \tag{9}$$

where *k* is the rate constant, *t* is time taken for the reaction, [*C*]t is the concentration of chlorpyrifos when time is equal to t, and [*C*]0 is the initial concentration of chlorpyrifos. The reaction rate is thus given by Equation (10):

$$rate = k[\mathbb{C}]^2\tag{10}$$

#### **4. Conclusions**

The Mn-doped WO3/SnS2 photocatalyst was successfully synthesized, resulting in a highly crystalline structure. Rectangular rods and sheet-like shapes were observed in the composite, confirming that no shape distortion had occurred in the heterojunction photocatalyst. The composite comprises both hexagonal and monoclinic phases that correspond to SnS2 and WO3, respectively, as confirmed by XRD patterns and Raman spectra. As shown in the UV-Vis spectra of the composite, a shift in the *band edge* (*absorption band edge*) from the UV to the visible region (red shift) was observed in the Mn-doped WO3/SnS2 photocatalyst relative to that for the pristine photocatalysts. The surface area of the WO3 was improved by more than 10 times by intrinsic doping with the Mn2<sup>+</sup> ion and the formation of the heterojunction with SnS2 to form the Mn-doped WO3/SnS2 photocatalyst. The Mn-doped composite was fully characterised using microscopic and spectroscopic techniques, which confirmed the synthesised composite to be Mn-WO3/SnS2. The Mn-doped WO3/SnS2 showed good electrochemical performance, ascribed to its high current density and lower interfacial charge transfer resistance, observed using electrochemical measurements (EIS), which correspond to high charge separation and a low photogenerated charge carrier recombination rate, observed using photoluminescence (PL) measurements. Chlorpyrifos has been applied extensively in agriculture, both in South Africa and other parts of the world, to fight against pests, therefore finding its way into water systems. Chlorpyrifos removal from synthetic water was investigated using Mn-WO3/SnS2 nanoparticles. The removal was due to the enhanced charge separation, high charge transfers and high electrostatic attraction between the nanoparticles and chlorpyrifos.

After the optimization of the reaction conditions, the chlorpyrifos removal achieved was 99.99% at pH 7 with 1 g of Mn-WO3/SnS2 and a 1000 ppb concentration.

The degradation pathway was also investigated, for which 3,5,6-trichloropyridin-2-ol and O,O-dihydrogen phosphorothioite were observed. Furthermore, after 60 min of the reaction, only O,O-dihydrogen phosphorothioite was detected. This implies that both chlorpyrifos and TCP were completely degraded. The results suggest that our material, Mn-WO3/SnS2, can completely degrade chlorpyrifos and its major degradation product.

**Supplementary Materials:** The following are available online at http://www.mdpi.com/2073-4344/10/6/699/s1, Table S1: Average crystallite sizes of nanomaterials; Figure S2: (a) FESEM image of pristine Mn-WO3, (b) TEM image of Mn-WO3, (c–e) elemental mapping, and (f) EDX spectrum of Mn-WO3 nanoparticles; Figure S3: (a) FESEM image of pristine SnS2, (b) TEM image of SnS2 (inset is the corresponding SAED image), (c,d) elemental mapping, and (e) EDX spectrum of SnS2 nanoparticles; Figure S4: Absorption spectra of WO3, SnS2, Mn-WO3, WO3/SnS2, and Mn-WO3/SnS2; Figure S5: EIS spectra showing the fitted spectra when obtaining the Randles circuit for (a) WO3 and Mn-WO3, and (b) SnS2, WO3/SnS2 and Mn-WO3/SnS2; Figure S6: (a–e) N2 adsorption-desorption isotherm of (a) WO3, (b) Mn-WO3, (c) Mn-WO3/SnS2, (d) WO3/SnS2, and (e) SnS2 (insets are pore volume graphs); Figure S7: Calibration curve of chlorpyrifos standards from 3.125 to 75 ppb; Figure S8: Mass spectra showing m/z ratios from 0 to 60 min; Figure S9: Fitted second order reaction kinetics graphs of the nanoparticles.

**Author Contributions:** Conceptualization, L.N.D., methodology, C.M.K.; validation, L.N.D., S.P.M. and C.M.K.; formal analysis, C.M.K.; investigation, C.M.K.; resources, L.N.D.; data curation, C.M.K.; writing—original draft preparation, C.M.K.; writing—review and editing, L.N.D., S.P.M.; supervision, L.N.D., S.P.M.; project administration, L.N.D.; funding acquisition, L.N.D. All authors have read and agree to the published version of the manuscript.

**Funding:** This research was funded by THUTHUKA NATIONAL RESEARCH FOUNDATION, grant number 15060-9119-027" and "The APC was funded by UNIVERSITY OF JOHANNESBURG-Accelerated Academic Mentorship Programme".

**Acknowledgments:** The authors would like to extend their gratitude to the University of Johannesburg, Faculty of Science, National Research Foundation (NRF) (TTK 15060-9119-027), TESP Eskom and the Centre for Nanomaterials Science Research, University of Johannesburg.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


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