Degradation of Anti-Inflammatory Drugs in Synthetic Wastewater by Solar Photocatalysis

Abstract

Due to the high number of anti-inflammatory drugs (AIMDs) used by the public health sector in Iraq and distributed all over the country and due to their toxicity, there is a need for an environmental-friendly technique to degrade any wasted (AIMD) present in aquatic ecosystem. The degradation of diclofenac sodium (DCF), ibuprofen (IBN), and mefenamic acid (MFA) in synthetic hospital wastewater were investigated utilizing locally-made Cu-coated TiO₂ nanoparticles in a solar-irradiated reactor. Different key variables were studied for their effects on process efficiency, such as loadings of catalyst (C _CU-TiO2 = 100–500 mg/L), AIMDs (100 µg/L), pH (4–9), and hydrogen peroxide (C_H2O2 = 200–800 mg/L). The results revealed that degradation percentages of 96.5, 94.2, and 82.3%, were obtained for DCF, IBN, and MFA, respectively, using our Cu-coated TiO₂ catalyst within 65 min at pH = 9, while other parameters were C _CU-TiO2 = 300 mg/L, and C_H2O2 = 400 mg/L. The experimental results revealed coupling photocatalysis with solar irradiation as a clean energy source could be utilized for the degradation of toxic pollutants in surface water.

1. Introduction

Diclofenac (DIC), ibuprofen (IBN), and mefenamic acid (MFA) are nonsteroidal anti-inflammatory and analgesic drugs [1]. The chemical structures of these drugs are shown in Figure 1. Large quantity of anti-inflammatory drugs (AIMDs) are utilized by the health sector in Iraq and dispensed all over Iraq. Moreover, there are no clear policies for management of pharmaceuticals in the environment [2]. Wastewater discharged from some Iraqi hospitals comprises components of drugs which will enter the aquatic ecosystem if they are not removed. Toxicological studies have revealed that exposure to AIMDs can stimulate adverse effects on living aquatic organisms, for instance fishes may suffer kidney harm and changes in their gills if exposed to diclofenac at a loading of >1 µg/L for approximately 21 days [3]. Mehinto et al. [4] and Daouk et al. [5] also reported a syndrome of changes of swimming efficacy followed by death in organisms because of exposure to AIMDs.

The various water treatment systems used in pharmaceutical removal are particularly technologies using carbon adsorption, advanced oxidation (AOP) processes that use ultraviolet radiation, electro-oxidation, and membrane filtration methods such as nano- filtration and reverse osmosis [6].

The advanced oxidation process (e.g., heterogeneous photocatalysis) encompasses the use of a solid semiconductor such as ZnO and TiO₂ to form an irradiation-effect stable suspension, to enhance the chemical reaction into the operating system. The best semiconductor in heterogeneous treatment is recently considered TiO₂. Read more about this TiO₂. The most practicable deployment of AOPs as shown by a wide range of literatures relating to the elimination of toxic organic products from water is the decay of organic compounds by the heterogeneous photoinduced reaction (HPR) [7].

Figure 2 summarizes the kinetic mechanism of AOP. Electron-hole pairs (e⁻-h⁺) pairs are formed when TiO₂ surfaces are irradiated by light photons having energy greater than 3.2 eV. These pairs generate hydroxyl radicals (^•OH) when combined with H₂O molecules and dissolved oxygen (DO). (H⁺) ions, superoxide radical (O₂^−•), and peroxide radicals (^•OOH) are also formed. (^•OOH) joins with H⁺ to produce ^•OH and OH⁻. (h⁺) oxidizes OH⁻ to ^•OH. Hence, all species finally enhance the generation of ^•OHs. These radicals break up different toxic chains that exist in the liquid [7,8].

It is well-known that TiO₂ utilizes a UV bandwidth to initiate a photoinduced reaction at 3.2 eV. Copper oxides display an enhanced absorptive material with a narrow band gap (1.4–2.2 eV). Copper oxides in conjunction with TiO₂ were stated to give greater stability and greater efficacy photocatalytic [9]. Many investigations investigated the deterioration of toxic material produced from industrial activities (i.e., xylene, antibiotics, phenol, and pesticides) through the use of homogenous photocatalytic processes using a constant UV or variable solar radiating reactor with a variety of operational parameters: the pH of the solution, the catalyst concentration, loading of a homogeneous catalyst, H₂O₂ loading, and liquid flow rate [10,11,12,13,14]. In recent studies, Khan et al. [15] investigated the photocatalytic performance and the biocidal potential of a synthesized nitrogen-doped TiO₂ nanoparticles against reactive black 5 (RB5), a double azo dye and E. coli. The authors reported that for the RB5 90% COD removal was achieved after 60 min of irradiation, confirmed by the disappearance of spectral peaks while the best-optimized photocatalysts showed a noticeable biocidal potential against human pathogenic strain E. coli in 150 min. Moreover, Khan et al. [16] studied the photocatalytic activity of a prepared graphene quantum dots (GQDs) and iron co-doped TiO₂ nanoparticles for decolorization of reactive black 5 dye. The authors revealed that GQD-0.1Fe co-doping of TiO₂ greatly improved the photocatalytic decolorization efficiency for RB5 dye.

The current research aimed to study AIMDs degradation in synthetic wastewater using a solar-irradiated catalytic reactor, in accordance with various operating parameters as solar irradiation intensity, solution acidity, H₂O₂, and a photocatalyst concentration in the solution.

2. Materials and Methods

2.1. Materials

Ti{OCH(CH₃)₂}₄ (purity 98%) was supplied by Chempure (Mumbai, India). Alfa Aesar (Jakarta Utara, Indonesia) provided the CuSO₄·5H₂O (98% min). Sigma-Aldrich (Bangalore, India) provided with analytical grade samples of diclofenac (DIC), ibuprofen (IBN) and mefenamic acid (MFA). H₂O₂ (30%) was supplied by Arkema (Istanbul, Turkey). The local market supplied RO water, HCl (36%), NaOH (99%), and NaCl (99%), while Sigma-Aldrich (OMA International Business, Baghdad, Iraq) supplied analytical grades of potassium chloride, potassium dihydrogen orthophosphate (KH₂PO₃) and disodium hydrogen orthphosphate (Na₂HPO₄).

2.2. Methods

2.2.1. Catalyst Preparation

The Cu-coated TiO₂ nanoparticle was synthesized following the procedure of [9]. One M CuSO₄ was prepared by dissolving CuSO₄·5H2O in NH₄OH (25 wt%). One mM (97% min) of Ti{OCH(CH₃)₂}₄ was blended with the 1 M CuSO₄ solution and the blend was sonicated for half an hour, with 5 min sonication intervals, followed by a rest period of 3 min, using an Ultrasonic model system (VCX-750 Vibracell, SONICE, Newtown, CT, USA). The sediment obtained by filtration with double 1 µm filter papers, was washed with acetone and ethanol several times to remove any remaining organic matter. For 1 h at 110 °C, the sediment was dried. The calcination of the as-synthesized Cu-doped TiO₂ was done for 60 min at 500 ± 2 °C [9]. Cu-free TiO₂ was synthesized following the same steps but without using CuSO₄.

2.2.2. Catalyst Characterization

The surface morphology was examined using an Inspect S50 SEM system (FEI, Eindhoven, The Netherlands) with voltage of 200 V to 30 kV and a magnification range from 4× to 105×. FTIR was used to determine the various functional groups present in the samples [17]. The FTIR spectra were obtained in the 400–4000 cm⁻¹ range using a FT/IR-6300 spectrometer (Jasco, Easton, MD, USA). The pHs of solutions were monitored by a pH 9124 device made by HANNA (Woonsocket, RI, USA) equipped with Ag/AgCl electrodes.

2.3. Experimental Setup

The experiments were conducted using a solar reactor with a usable volume (reactor volume + tank volume) of 30 L. A batch mode was used in the solar reactor. The frame includes panels, wastewater synthesis vessels, a centrifugal pump and a control panel. The sunlight collector was placed on a 37° (regional latitude) steel base corresponding to the solar orientation of the southern city of Baghdad (Iraq). Ten cylindrical Pyrex glass tubes of 1.25 m in length and 2.5 cm in diameter, respectively, were used for the solar collectors. The pipes were connected by custom PVC fittings. A balanced flow meter was used to load wastewater from the vessel to the collectors. Polished panels of aluminum were arranged as a tub shape and were placed on the stainless-steel stand. This design requires light to be reflected into the central line of pipes from all sides, and even for photocatalyst the light entering pipes can be used. A temperature sensor was mounted at the reactor exit to measure the temperature, and samples are taken periodically to determine the concentration. To maintain a uniform tank level and maintain aeration, a mechanical stirrer was used. The entire configuration was managed by a control panel. Figure 3 gives an overview of the photocatalytic system setup. It is worth noting that before starting irradiation with solar light, experiments of AIMDs adsorption onto TiO₂ were conducted in the dark. The suspensions were circulated through the system during 2 h in the absence of light in order to reach equilibrium adsorption in the presence of oxygen obtained from the air. The “initial” concentration C₀ is therefore slightly lower than the original concentration of the solution due to a partial adsorption (of about 5%). Preparation details of buffer solution and calibration curves are given in Supplementary Material.

2.4. Range of Operating Parameters

Table 1. Operating variables.

AIMD Conc. µg/L	pH	H2O2 Loading mg/L	TiO2 Loading mg/L
1	4	200	100
30	7	400	300
60	9	500	500
100	9	500	500

shows the operating variables with their ranges. In all the experiments the wastewater flow rate was kept at 1 L min⁻¹.

2.5. Solution Preparation

To make a buffer of pH 4, 0.1 M KCl solution was poured slowly to 0.1 M HCl till the desired pH was obtained. For the synthesis of pH 7 buffer, KH₂PO₄ (0.6 g), Na₂PO₄H (6.4 g) and NaCl (5.85 g) were dissolved in an excess of distilled water to yield 1000 mL and the pH was regulated with a 15% NH₄OH. Similarly, pH 9 buffer was made by dissolving KH₂PO₄ (17.4 g) in distilled water (800 mL) and the pH was adjusted with a 15% NH₄OH solution.

To establish the calibration curve, working standard solutions of diclofenac sodium, ibuprofen, and mefenamic acid in concentrations ranging from 1 to 120 µg L⁻¹ were prepared in pH 4 buffer. For these solutions, the maximum absorbance was screened against the reagent blank in the region from 190 to 700 nm. This process was then replicated using buffers of pH 7 and 9. The calibration curves of diclofenac sodium, ibuprofen, and mefenamic acid at pH 9 can be seen in Figure S1 in the Supplementary File.

3. Results and Discussion

3.1. Surface Morphology and Cu-Coated TiO₂ Catalyst Structure

3.1.1. SEM Analysis

Scanning electron microscopy (SEM) is widely used in science to characterize the surface roughness of materials [18,19] because from SEM images the actual particle size can be easily determined. In Figure 4a (SEM Mag. 1 kx) and Figure 4b (SEM Mag. 2.00 kx), SEM photographs of 3 wt% of Cu-doped TiO₂ are presented. The SEM images show a few localized agglomerations with a pattern of cracks. A variety of 57 and 84 nm nanocrystallites along with nano-sized TiO₂ are seen in the images. That is, a decrease in particle size due to Cu coating is observed. When Cu²⁺ ions are doped into TiO₂ lattice, they like to locate in grain boundary regions or on the surface of particles where they inhibit the growth of TiO₂ crystals.

3.1.2. EDS Analysis

EDS can be used to determine which chemical elements are present in a sample [20,21]. Pure TiO₂ and the Cu-doped TiO₂ catalyst exhibited distinctive surface characteristics in the EDS plots, as illustrated in Figure 5. The horizontal coordinate of the EDS picture represents the energy of ionization, and the vertical coordinate shows the count. The higher the count, the more of that component was present in the observation area. Comparing uncoated TiO₂ with the TiO₂ coated with 3.5 wt% Cu (Figure 5b) reveals peaks of 3.5 wt% Cu-doped TiO₂ and the presence of Cu metal with a high content of TiO₂ atoms, as displayed in Figure 5a. Thus, the Cu presence in the TiO₂ grid was obvious. The results are seen in Figure 5a, which depicts a percentage of TiO₂ and oxygen of 62.5 and 37.5 wt% respectively, while in Figure 5b, the EDS spectra show percentages of TiO₂, oxygen, and Cu of 47.5, 49, and 3.5 wt%, respectively.

3.2. FTIR Analysis

The FTIR spectra for TiO₂ and Cu- coated TiO₂ are seen in Figure 6a,b, respectively. The FTIR spectrum of TiO₂ has two main bands, one wide and extreme with a peak of 3420 cm⁻¹, the other at 1620 cm⁻¹. The peaks at 3420 and 1620 cm⁻¹ can be due to O–H moieties from moisture hydroxyl groupings and bending that occurs on the catalyst surface. For the Cu-doped TiO₂ spectrum the O–H group is shown at the peaks of 3460 and 1722 cm⁻¹. Vibrations of Ti-O and Ti-O–Ti framework bands of TiO₂ can be attributable to the peaks of TiO₂, at 476, and 620 cm⁻¹ [21]. There was no characteristic Cu-O peak in the area between 400 and 600 cm⁻¹ (432.3, 497 and 603 cm⁻¹) [22,23]. Since the volume of metal salts is minimal, there is no new band and no change in the nano-TiO₂ bands after Cu doping.

3.3. Operating Parameters Effect on DIC, IBN, and MFA Degradation

All figures were plotted to measure AIMDs’ degradation as a function of incident solar energy. The accumulated energy per unit volume of synthetic wastewater flow (${\mathrm{Q}}_{\mathrm{UV},\mathrm{n}}$) across the solar reactor of the n-th sample was estimated using Equation (1) [24]. The solar light intensity was measured by using Davis instruments 6152C Vantage Pro2 Weather Station (USA):

$${\mathrm{Q}}_{\mathrm{UV},\mathrm{n}}={ \mathrm{Q}}_{\mathrm{UV},\mathrm{n}-1}+\Delta {\mathrm{t}}_{\mathrm{n}}{ \mathrm{UV}}_{\mathrm{GN}} \frac{\mathrm{A}}{\mathrm{V}}$$

(1)

where ∆t is the time interval between each sample (15 min), A is the irradiated area of the reactor (10 × π × 2.5 × 125) cm², V is the volume of synthetic wastewater in the system = 30 L, n is the number of samples, and UV_GN: is the average local global UV irradiation = 33.7 W/m².

3.3.1. Influence of pH

Figure 7a–c illustrate the removal efficiency of AIMDs versus the accumulated solar energy per unit volume of wastewater (Q_UVS) at the reactor outlet at various pH values after 65 min. The pH was changed from 4, 7, and 9 keeping other variables constant (C_MFA = 30 µg/L, C _CU-TiO2 = 300 mg/L, and C_H2O2 = 400 mg/L).

Figure 7a shows that after 65 min of irradiation, the removal efficiency of MFA was 40.0, 50. 0, and 82.3 at pH 4, 7, and 9, respectively. Figure 7b illustrates that after 65 min of irradiation, the removal efficiency of DCF was 50.0, 73. 0, and 96.5 at pH of 4, 7, and 9, respectively. The removal efficiency of IBN is seen in Figure 7c. As can be observed in this figure the %R of IBN was 48.0, 72.0, and 94.3 at pH of 4, 7, and 9, respectively. This behaviour could be imputed to the changes on the TiO₂ surface.

In an alkaline medium, TiO₂ acquires a charge that creates electric drag forces on the TiO₂ surface and the AIMD ions in the wastewater enhancing the sorption of these ions on the illuminated surface of dopedTiO₂ and hence increasing the AIMD removal. However, in the acidic environment, a different effect was observed. This may be attributed to the repulsive electrostatic forces between the nanoparticles of TiO₂ and AIMDs ions in the solution hindering the adsorption over the Cu-doped TiO₂. Poulios et al. introduced Equations (2) and (3) describing the alteration in the surface behavior with changing pH of the solution near its pHpzc [25].

$$\mathrm{TiOH}+{ \mathrm{H}}^{+}\to {\mathrm{TiOH}}_2^{+}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em} \mathrm{pH} < {\mathrm{pH}}_{\mathrm{pzs}}$$

(2)

$$\mathrm{TiOH}+{ \mathrm{OH}}^{-}\to {\mathrm{TiO}}^{-}+{ \mathrm{H}}_2\mathrm{O}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em} \mathrm{pH} > {\mathrm{pH}}_{\mathrm{pzs}}$$

(3)

As a consequence, the aforementioned pH alterations influence the sorption of AIMDs ions over surface of TiO₂, a basic step for the photo-oxidation to take place. In our present work the pHpzc of Cu-doped TiO₂ photocatalyst was measured using the electrolyte titration method of Preocanin and Nikola [26]. The pHpzc ranged from 5.9–6.0 (see Figure 8), and when the pH of suspension exceeds 6.0 the number of sorbed AIMDs ions over the TiO₂ begins to grow because of the increasing of TiO⁻ presence in the TiO₂. In our work, the percentage removal of all AIMDs reached a maximum at a pH of 9. Because of this, the removal of AIMDs attains high levels in alkaline solutions [24].

3.3.2. Effect of H₂O₂ Loading

Figure 9a–c depict the effect of different H₂O₂ loadings (200 to 500 mg/L) on the removal efficiency of AIMDs under solar illumination, when other operating variables remain unchanged. As can be seen in Figure 9a, as the loading of H₂O₂ increased from 200, to 400, and 500 mg/L the %R of DCF becomes 48.0, 96.5, and 90.0%, respectively. In Figure 9b, the removal efficiency of MFA after 65 min of solar light irradiation with the same additions of H₂O₂ is shown to be 47, 76, and 82.3%, respectively, whilst in Figure 9c, as the loading of H₂O₂ increased from 200, 400, and 500 mg/L the %R of IBN becomes 60.0, 94.3, and 88.0%, respectively. In all figures, it is observed that as the H₂O₂ concentration was raised, %R for AIMDs exhibited the same behavior attaining the highest value at 400 mg H₂O₂/L, and after this %R started to decrease with further addition of H₂O₂. This may suggest that the optimum concentration of H₂O₂ for the best removal efficiency is 400 mg/L which was employed in all experiments. The mentioned trend may be due to that fact that at 200 mg/L loading an inadequate amount of hydroxyl radicals was generated, which are responsible for AIMDs’ removal. Moreover, H₂O₂ hinders the recombination of the electrons and holes over the surface of TiO₂ and increases the generation of ^•OH to the photocatalyst [26]. However, as the H₂O₂ loading increased more than 400 mg/L, the rate of increase in %R was less. This may be attributed to that as more quantity of H₂O₂ is added more ^•OH radicals are generated, yet the excess of H₂O₂ react with the hydroxyl radicals generated rather than with the organic substrate, according to Equations (4) and (5) [11,24]:

H₂O₂ + ^•OH → H₂O + HO₂

(4)

HO₂ + ^•OH → H₂O + O₂

(5)

Therefore, it is imperative to determine the stoichiometric amount of hydrogen peroxide sufficient for complete mineralization. H₂O₂ appears to have two functions in the reaction of photooxidation, firstly it hinders electron-hole recombination and it can also split to form ^•OH [27].

3.3.3. Effect of Catalyst Loading

Changes of %R with solar light for various TiO₂ loading (100–500 mg L⁻¹) maintaining other process parameters unchanged are seen in Figure 10a–c for DCF, IBN, and MFA, respectively, after 65 min of illumination. In Figure 10a it is observed that as the TiO₂ loading increased from 100, 300, and 500 mg L⁻¹ the %R of DCF increases from 55, 96.5, and 78%, while in Figure 10b it is shown that as the TiO₂ loading increased from 100, 300, and 500 mg L⁻¹ %R of IBN increased to 52, 94.3, and 68%. In Figure 10c it is seen that as the TiO₂ loading increased from 100, 300, and 500 mg L⁻¹, the %R of MFA increased 37, 82.3, and 54%.

In all plots of Figure 10 it is noticed that with each specified quantity of TiO₂ concentration in suspension the removal efficiency of AIMDs increased steeply with the Q_UV till it reached a certain magnitude beyond which the rate growth became limited. This pattern can be due to the reaction rate of the AIMDs, which is slower when the amount of MFA used is smaller [28,29]. Figure 10 also shows the effect of the average rise in TiO₂ concentration on the removal efficiency; this trend is attributed to the increase in TiO₂-active sites with more surface-adsorbed cations of AIMD. In order to prevent the use of excess TiO₂, the most favorable load for the efficient removal of AIMDs must be predicted. Figure 10 shows that the rise in the TiO₂ concentration by R is consequently increased from 100–300 mg L⁻¹. This may be attributed to the number of active catalyst sites which increased enhancing the photoreaction. On the other hand, a different trend was observed for TiO₂ concentrations of more than 300 mg L⁻¹. This trend occurred due to the reduction in light permeation into the wastewater due to the rising density of the suspension eventually becomes a lightproof medium [25,30].

3.4. Comparison of Catalysts Performance

The comparison between Cu coated TiO₂ and pure TiO₂ for AIMDs degradation using the optimum parameters is illustrated in Figure 11, where 3.5 wt% Cu-coated TiO₂ exhibits a noteworthy growth in the mefenamic acid degradation performance under solar light. The results depicted that about a 51% raise in MFA removal was noticed for the 3.5 wt% Cu-coated TiO₂ over that achieved over pure TiO₂. This could be due to the influence of the Cu powder distributed over the TiO₂ nanoparticles causing a decrease in the bandgap of TiO₂ from 3.2 to 2.2 eV allowing the participation of visible light irradiation in addition to UV in the photodegradation mechanism. Moreover, about a 1% raise in MFA removal was noticed for the 3.5 wt% Cu-coated TiO₂ over that of 2.5 wt% Cu-coated TiO₂ which was synthesized by Shanian et al. [24]. One may conclude that the presence of the optimum stoichiometric amounts of titanium dioxide with Cu is necessary to display a reasonable photocatalytic activity. Our outcome displayed the same trend as that found by Czupryn et al., [31] who synthesized Pt/TiO₂ to oxidize gaseous CO under UV light. The present results also revealed that the optimum value of Cu coating TiO₂ was 2.5 wt% from the cost point of view.

To validate these results a light absorbency measurement has been conducted for the synthesized photocatalysts as shown below. Cu-coated TiO₂ nanoparticles with different wt% of Cu (0, and 3.5%) were synthesized and UV-VIS absorbance of doped and undoped TiO₂ was measured using a Shimadzu UV- Vis 1601 spectrophotometer. The bandgap energy of the nano photocatalyst is calculated according to the formula $E=\frac{h c}{\mathsf{\lambda }}$ [32]. where h, c, and $\lambda$ are Blank constant (=4.1357 × 10⁻¹⁵ eV·s), light velocity (=2.99705 × 10⁸ m·s⁻¹), and wavelength (nm) respectively. Absorbance data for the samples are listed in

Table 2. Effect of Cu coating on the band gap energy of the photocatalyst.

Item	wt% Cu-Coating	Wavelength (nm)	(eV)	Ref.
1	0	3.87	3.20	present work
2	2.5	410	2.20	[24]
2	3.5	415	1.98	present work

. The table shows a shift in wavelength from the UV to the visible region. The absorbance values start increasing as the Cu loading is increased.

4. Conclusions

Experimental studies using solar irradiation of removing anti-inflammatory drugs (AIMDs) from wastewater have shown that different operational variables, for example the acidity of the wastewater, photocatalyst, and hydrogen peroxide loadings all affect the removal effectiveness. The optimum operating conditions to obtain the best photodegradation of the tested AIMDs were pH = 9, C_(Cu-TiO2) = 300 mg/L, and C (H₂O₂) = 400 mg/L. The increased loading of H₂O₂ above the optimal value has no impact on the elimination of AIMDs. Moreover, as the photocatalyst concentration was increased above an optimum, solar light penetration through the liquid film decreased, reducing the efficiency of the photodegradation. Cu-coated TiO₂ catalyst has a strong effect on improving the elimination of AIMD. Operating with 3.5 wt% Cu-coated TiO₂ increased the elimination of AIMDs by around 51% over that of pure TiO₂ since the bandgap of pure TiO₂ was decreased from 3.2 to 2.2 eV by the coating procedure. However, it has been found that increasing Cu coating over 2.5 wt% has only a slight effect on the degradation of AIMDs. The findings of this research revealed that solar irradiation could be used as a clean energy source for the degradation of AIMDs in surface waters.