*Article* **Photo-Catalytic Remediation of Pesticides in Wastewater Using UV/TiO<sup>2</sup>**

**Mohamed H. EL-Saeid 1,\*, Modhi O. Alotaibi 2,\* , Mashael Alshabanat 3,\*, Khadiga Alharbi <sup>2</sup> , Abeer S. Altowyan <sup>4</sup> and Murefah Al-Anazy <sup>3</sup>**


**Abstract:** One of the most serious environmental concerns worldwide is the consequences of industrial wastes and agricultural usage leading to pesticide residues in water. At present, a wide range of pesticides are used directly to control pests and diseases. However, environmental damage is expected even at their low concentration because they are sustained a long time in nature, which has a negative impact on human health. In this study, photolysis and photocatalysis of the pesticides dieldrin and deltamethrin were tested at two UV wavelengths (254 and 306 nm) and in different test media (distilled water, wastewater, and agricultural wastewater) to examine their ability to eliminate pesticides. TiO<sup>2</sup> (0.001 g/10 mL) was used as a catalyst for each treatment. The purpose was to determine the influence of UV wavelength, exposure time, and catalyst addition on the pesticide decomposition processes in different water types. Water was loaded with the tested pesticides (2000 µg) for 12 h under UV irradiation, and the pesticide concentrations were measured at 2 h intervals after UV irradiation. The results showed a clear effect of UV light on the pesticides photodegradations that was both a wavelength- and time-dependent effect. Photolysis was more effective at λ = 306 nm than at λ = 254 nm. Furthermore, TiO<sup>2</sup> addition (0.001 g/10 mL) increased the degradation at both tested wavelengths and hence could be considered a potential catalyst for both pesticide degradations. Deltamethrin was more sensitive to UV light than dieldrin under all conditions.

**Keywords:** photolysis; catalysis; degradation; pesticides; UV; wastewater; agricultural wastewater

#### **1. Introduction**

Agrochemicals are substances that are commonly used in agriculture to protect crops and ensure their productivity [1]. Such chemicals are commonly applied to eliminate pests (such as rodents), and include pesticides—namely, insecticides, fungicides, and herbicides—and un-wanted plants [2]. In public health, agrochemicals are used to combat human disease vectors such as mosquitoes; they are also used against crop-damaging epidemics in the agricultural sector [3,4] that also offer producers an efficient means to manage crop pests that decrease yield and threaten food security [5]. While some are used at a crop's initial production stage, others are generally used on edible plant parts before harvest or even during storage. Therefore, crop-based agrochemicals such as pesticides are dissolved in water, and the crop are sprayed in the fields.

Many new pesticides have been introduced over the last few decades, which have toxic effect in the short and/or long term [3,4]. Commercial pesticide formulations often include additional compounds (such as solvents and surfactants) to increase their

**Citation:** EL-Saeid, M.H.; Alotaibi, M.O.; Alshabanat, M.; Alharbi, K.; Altowyan, A.S.; Al-Anazy, M. Photo-Catalytic Remediation of Pesticides in Wastewater Using UV/TiO2. *Water* **2021**, *13*, 3080. https://doi.org/10.3390/w13213080

Academic Editor: Chengyun Zhou

Received: 28 July 2021 Accepted: 21 October 2021 Published: 2 November 2021

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

**Copyright:** © 2021 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 (https:// creativecommons.org/licenses/by/ 4.0/).

activity; solvents and other co-formulants often increase the environmental impact of the formulation as well.

The potential harmful effects of used pesticide on both humans and the environment have received growing attention from the community and expert authorities. Numerous studies have focused on health or environmental concerns from accidental or intentional pesticide exposure, specifically those highly toxic to mammals or found in the environment. The pesticide's risks should be reduced to their minimum via careful regulation and appropriate user guidance. However, the positive effects of pesticide use should not be overlooked. When rational, careful use of pesticides in combination with other technologies is considered in integrated pest management systems, their usage is likely to be justified [6].

On the other hand, fresh fruit and vegetable growers use various water sources, including surface water sources (such as rivers or lakes) that are potentially contaminated by chemical pollutants [7]. Thus, the water used in agricultural production increasingly might have the potential to introduce pathogenic viruses into fresh produce supply chains. The Codex Committee on Pesticide Residues' Code of Hygienic Practice for Fresh Fruits and Vegetables (adopted in 2003; revision in 2010 (new Annex III for Fresh Leafy 334 Vegetables), from the Netherlands, March 2010 [7], pointed out the importance of using 'clean water' for fresh produce cultivation, particularly if water is applied before harvest and in close contact with the edible plant part. Environmental protection programs seek to encourage a reduction in the use of pesticides as a precaution product in growing crops [8].

One of the most abundant water contaminants is pesticide residue, and despite pesticides' economic advantages, such as high crop yields, their potential health hazards are still unknown. Water bodies are polluted by toxic chemicals resulting from human activities in industry, agriculture, and housing [9]. The residues of industrial and agricultural areas contaminated with pesticides are dumped unattended into the nearby water bodies, although most of them are not degradable in water, and thus the aquatic environment becomes threatened [10]. The effects of pollutants are generally characterized by alteration in the animal physiological behavior, and therefore affect survival, reproduction, and growth.

Most Europeans are concerned primarily with the long-term or chronic consequences of low exposure levels through various pathways, especially residues in food crops, as well as through pesticide fraction losses from the target areas [11].

Dieldrin and deltamethrin are environmental pollutants with long-term adverse effects. The use of dieldrin has been banned in most countries around the world primarily for environmental reasons. The widespread use of dieldrin and its ecological persistence have resulted in survival in the environment [12], with bioaccumulation in the food chain due to their low volatile, chemically stable, and lipophilic properties [13]. The half-life of dieldrin is about 5 years [14]. In addition, it will take 25 years for 90% to disappear [15] and remains for 60 years when not exposed to sunlight. In addition, it was used for termite control until about 1985, which means it is still in the basement of the most houses right now and in the soils of agricultural fields with these pesticides [16]. In a previous study published in 2005 [17], Saqib et al. identified residues of DDT, DDE, aldrin, dieldrin, and deltamethrin in fish tissues in Haleji Lake, and more different pesticide compounds were identified in Kalri Lake, possibly because of runoff from surrounding agricultural farms [17]. On the other hand, in Saudi Arabia at Al-Qassim, high pesticide concentrations might be linked to intensive agricultural activity [18]. Researchers [19] stated that dieldrin and deltamethrin are pesticides found in fruits and palm in Riyadh market, Saudi Arabia, which might suggest contamination of irrigation water. The search for a potential tool such as photolysis for the above-mentioned pesticide residue degradation is important.

Recent studies are starting to shed light on disposal strategies for pesticides in contaminated water that would lead to better effluent water quality and have focused on the possibility of analyzing pollutants and removing pollution by available and less expensive methods [20–25]. Currently, several processes have been developed to reduce harmful pollutants in wastewater, including advanced oxidation processes (AOP) [26], activated sludge treatments [27], electro-removal [28], ozonation [29], sunlight [30], UV radiation [31], and combined/integrated methods [32]. Catalyst methodologies have also been used to improve the pesticide disposal mechanism. Photodegradation is induced by the action of light and is attributed to chemical reactions arising by photoionization. Researchers [33] explained that one of the most important abiotic transformations of pesticides in the aquatic environment is photolysis, where the high energy of solar rays causes characteristic reactions such as bond separation, rotation, and rearrangement. Photolysis with the aid of catalytic compounds could be beneficial when UV radiation is applied. UV radiation has adequate energy for chemical bond breakdown; the high-energy photons cause ionization.

Dieldrin and deltamethrin are environmental pollutants that are prohibited in most countries around the world, but they are still used, leading to contamination of many environments, such as soil, sediment, and groundwater [16,34]. Therefore, looking for solution to reduce their negative impact is an urgent issue. UV has been well known as water disinfectant for microbial removal [35,36], and as an efficient technique for the treatment of wastewater [37]. However, new studies to develop the photo-remediation by UV radiation as an effective method for water treatment systems are needed.

The purpose of this work was to obtain information that is currently lacking regarding the photo-remediation of the pesticides dieldrin and deltamethrin, since their high levels in contaminated water are expected and attributed to pollution as consequences to industrial wastes and agricultural usage. Therefore, in this study, different types of water contaminated with dieldrin and deltamethrin were irradiated at two UV wavelengths, with and without a catalyst, to observe the effect of the potential catalyst and identify the most effective UV wavelength and time span for maximum pesticide decomposition.

The importance of this study is in the remediation of dieldrin- and deltamethrinpolluted water taken from the local environment, including treated wastewater and agricultural wastewater. It will provide new data and potential breakthroughs to scientists, especially those working in environmental pollution and water remediation.

#### **2. Materials and Methods**

#### *2.1. Study Samples*

Three water samples were used as targets in the study: distilled water, wastewater, and agriculture wastewater. The distilled water used was obtained from a Millipore distilled water system (College of Food and Agricultural Sciences, King Saud University, Riyadh, Saudi Arabia).

First, 5 L of the wastewater sample was taken in dark glass container from treatment plants at Al Mansuriyah, Riyadh (pH = 7.33, EC = 1.77 µS/cm, TDS = 1133 mg/L, turbidity = 1.89 NTU) where the treatment process was performed by activated sludge method using tertiary treatment, and 5 L of the agriculture wastewater was taken from the Al-Kharj agricultural region (pH = 8.42, EC = 2.48 µS/cm, TDS = 1579 mg/L, turbidity = 3.24 NTU). The crops produced by the farms were corn, grains, dates, and some vegetables and leafy crops.

Both samples were transferred under cooling within 2 h to the analysis and experimental lab. Wastewater samples were analyzed for pesticide residues within 24 h and then the treatment of remediation began.

Pesticide residues in the collected wastewater samples were analyzed before spik-ing with 2000 ppb concentration of dieldrin and deltamethrin.

#### *2.2. Standards and Reagents*

Dieldrin, deltamethrin (Table 1 and Figure 1), calibration, and injection standards (99.9% purity) were purchased from AccuStandard, Inc., New Haven, CT, USA as individual or mixture standards at concentrations of 10 µg/mL. All internal standards were 13C 12-labelled (13C-labelled compound use allowed for the analysis to be quantified without clean up). All solvents used for the extraction and analysis of pesticides were analysis-grade residues (99.9% purity) and were obtained from Fisher Scientific (Fair Lawn, NJ, USA). QuEChERS kits were purchased from Phenomenex (Torrance, CA, USA). Titanium dioxide (TiO2) as a photocatalyst was from Sigma-Aldrich Chemie GmbH, Germany

(molecular weight: 79.87, CAS Number: 13463-67-7, 718467nanopowder, 21 nm primary particle size (TEM), ≥99.5% trace metals basis).


**Figure 1.** A structure of Dieldrin and deltamethrin.

#### *2.3. Sample Remediation by UV Photolysis (UV)*

The water was photo-treated using ultraviolet radiation at 254 and 306 nm wavelengths for the two pesticides' decomposition. Boekel UV Crosslinker (BUV) model 234100-2: 230 VAC, 175 W, 0.8 A was applied with four 254 nm lamps and Boekel Scientific, 855 Pennsylvania Blvd. Feasterville, PA, USA with four 306 nm lamps. The lamps and water samples were at a 15 cm distance at 1071 µWcm−<sup>2</sup> intensity of UV irradiation. Each pesticide (approximately 2000 µg/L) was loaded into the water and incubated for 12 h under UV lighting. Samples were taken for pesticide quantity residue measurement at 2 h intervals to identify the correct UV wavelength and the photolysis process exposure time. Furthermore, the same procedure was repeated for each pesticide with the addition of 0.001 g TiO<sup>2</sup> to each 10 mL water sample to study the effect of the catalyst.

#### *2.4. Samples Extraction and Cleanup by QuEChERS*

First, 10 mL of the water sample was transferred into a 50 mL centrifuge tube and vortexed briefly. After that, 10 mL acetonitrile was added to each sample and shaken using a vortex for 5 min to extract the pesticides, using a Spex Sample Prep Geno/Grinder 2010 operated at 1500 rpm. Next, the contents of an ECQUEU750CT-MP (citrate salts) Mylar pouch were added to each centrifuge tube. The samples were then shaken for at least 2 min and centrifuged for 5 min at ≥3500 rcf. A 1 mL aliquot of supernatant was transferred to a 2 mL CUMPSC18CT (MgSO4, PSA, C18) dSPE tube. The samples were shaken in a vortex for about 1 min, then centrifuged for 2 min at high rcf (e.g., ≥5000). The purified supernatant was filtered through a 0.2 µm syringe filter directly into a GC sample vial, and thereafter the sample was kept for further analysis.

#### *2.5. Analysis by Triple-Quadrupole Gas Chromatography Mass Spectrometry (GCMSMSTSQ 8000/SRM)*

The analysis was carried out using the latest Thermo Scientific™ TSQ 8000™ triplequadrupole GC-MS/MS system equipped with the Thermo Scientific™ TRACE™ 1310 GC with SSL Instant Connect™ SSL module and Thermo Scientific™ TriPlus™ RSH auto sampler (Waltham, MA, USA). The transition conditions are presented in Table 2.


#### **Table 2.** GCMSMSTQD 8000 SRM instrumental conditions.

#### *2.6. QAQC Strategies and Method Performance*

For quality analysis and quality control, samples were prepared in triplicate, blanked, and spiked. Certified reference material (CRM) was prepared and processed with each batch (5–10 samples) analyzed. QuEChERS and GCMSMSTSQ 8000/SRM method limit detection (LOD) and limit quantification (LQD), repeatability, reproducibility, accuracy, and precision were also determined for each pesticide (Table 3).

**Table 3.** Parameters of retention time, LOD, LOQ, recovery%, and GCMSTQD target mass of SRM scanning mode.


#### **3. Results and Discussion**

#### *3.1. Photolysis Process*

The photolysis process of dieldrin and deltamethrin was examined in the current study for three different types of water: distilled water (DW), wastewater (WW). and agricultural wastewater (Ag.WW) using UV radiation at varied wavelengths with and without a catalytic agent (TiO2). The results indicated that the amount of both dieldrin and deltamethrin decreased gradually with increasing time after photolysis. The pesticides' degradation rate and reduction (%) was calculated as the variation between the concentration after treatment in relation to that before treatments. The reduction percentage in the concentration of pesticide residues after degradation process for dieldrin and deltamethrin in DW reached 39.35% and 73.6% at 254 nm and 43.95% and 76.55% at 306 nm, respectively, after 12 h of treatment. As for WW, reduction percentages of 43.3% and 83.8% at 254 nm and 49.3% and 84.35% were recorded after 12 h for dieldrin and deltamethrin, respectively, at 306 nm. Furthermore, in Ag.WW, the percentages of reduction of dieldrin and deltamethrin after 12 h were 58.8% and 46.8% at 254 nm versus 52.9% and 37.3% at 306 nm, respectively. It was observed that the longer UV wavelength (306 nm) had a higher capacity for pesticide degradation compared to 254 nm. The dieldrin and deltamethrin amounts in DW, WW, and Ag.WW samples were decreased with increased UV exposure time as indicated in Figures 2–4, and therefore, a time-dependent reduction was noted.

**Figure 2.** Concentration of pesticides (µg/L) versus exposure time of UV radiation for photolysis process at 254 nm (**left**) and 306 nm (**right**) in distilled water with and without the catalytic agent.

**Figure 3.** Concentration of pesticides (µg/L) versus exposure time of UV radiation for photolysis process at 254 nm (**left**) and 306 nm (**right**) in wastewater with and without the catalytic agent.

**Figure 4.** Concentration of pesticides (µg/L) versus exposure time of UV radiation for photolysis process at 254 nm (**left**) and 306 nm (**right**) in agricultural water with and without catalytic agent.

Experimental results indicated the abilities of the two tested wavelengths to promote pesticide photolysis. At both wavelengths investigated, deltamethrin was more degradable than dieldrin. In DW and WW, both pesticide degradations were observed after 4 h treatment at 254 nm; after 4 h at 306 nm, deltamethrin degradation was faster than that of dieldrin. However, in Ag.WW, both pesticide quantities were reduced from 2000 µg/L after the first 2 h at both wavelengths, suggesting higher UV efficiency in Ag.WW. After 2 h, the dieldrin amount detected was the same at both wavelengths, and a higher degradation was observed for deltamethrin, especially at 306 nm.

The longer UV wavelength (306 nm) showed a higher capacity for pesticide degradation compared with 254 nm, which is consistent with previous findings [38,39]; other pesticides are known to be degraded under UV exposure, suggesting that UV reactor usage might be a suitable approach for pesticide photolysis [40,41]. For example, in a previous study, the photolysis rate of deltamethrin and bifenthrin, another pyrethroid, under UV irradiation at 237, 240, and 246 nm was investigated by Tariq et al. [42]. Their findings revealed that deltamethrin was highly degradable in a time-dependent manner when subjected to UV irradiation in organic solvents. In the absence of UV light, the organophosphorus pesticide degradation rate was insignificant, indicating the significant role of UV in pesticide degradation [43].

This trend indicates that longer wavelengths lead to faster degradation as compared to that at shorter wavelengths, especially for deltamethrin. The destructive effect of UV on molecular bonds is well known; therefore, UV exposure should lead to increased pesticide degradation as time increases. Increased UV irradiation time increases the formation of free radicals in water, potentially leading to decomposition pesticide poisoning [44,45]. Furthermore, the difference in the degradation levels between deltamethrin and dieldrin may be due to differences in their structures.

#### *3.2. Photocatalysis Process*

The photo-remediation was performed with photocatalyst to study the effect of adding a catalytic amount of TiO<sup>2</sup> on the pesticide photodegradations [41,44]. Photocatalytic pesticide residue breakdown by oxidation processes (AOP) is a modern approach that uses photons to degrade pesticides to H2O, CO2, and inorganic compounds with no side effects [46]. However, catalyst type is an important factor in pesticide photodegradation [47]. Titanium dioxide (TiO2) was used as a photocatalyst because it is effective in the decomposition of organic compounds and is more photochemically stable in water [48]. It is considered as a beneficial material for wastewater treatment because of its safe character; it is used in different applications, mainly in environmental remediation [49].

The catalytic effect of TiO<sup>2</sup> (0.001 g/10 mL) added to aqueous media was tested when dieldrin and deltamethrin were exposed to ultraviolet irradiation. The degradation of pesticides with different wavelengths and exposure times was observed. The results are displayed in Figures 2–4, indicating degradation in a time-dependent manner, as was noticed for degradation without catalysis for both tested pesticides.

When TiO<sup>2</sup> was applied in DW, UV treatment led to complete disappearance of pesticides at 306 nm at the end of treatment time; however, deltamethrin disappeared at 254 nm and only 49% of dieldrin was identified after 12 h.

In addition, after adding the catalyst, it was noticed that the deltamethrin pesticide was not detected in all samples of water media of DW, WW, and Ag.WW after 12 h at any of the tested wavelengths, but only 49.9%, 49.1%, and 40.5% of dieldrin were detected in water media of DW, WW, and Ag.WW, respectively, at 254 nm, as well as only 32.6% and 24.3% of dieldrin were detected in WW and Ag.WW, respectively, when 306 nm was tested; however, after 12 h, no pesticides residues were detected in DW.

Furthermore, it was observed that the percentage of pesticide degradation was higher when the catalyst was present compared to the previous experiments without the catalyst.

Hydrolysis levels in the presence of the catalyst were higher because the final concentrations of the pesticides were low compared to those after remediation without the catalyst. Hence, this catalyst has a role in improving the photolysis process. Thus, pesticides can be effectively destroyed by photocatalysis in the presence of TiO<sup>2</sup> suspensions. The photo-remediation at both wavelengths of UV rays, with and without the catalyst, of the pesticides in different aqueous media as a function of time are displayed in Figures 2–4.

Effect of the addition of the photocatalyst on the degradation process has been reported in previous studies. Degradation of the compounds azinphos methyl, azinphos ethyl, disulfoton, dimethoate, and fenthion was detected in TiO<sup>2</sup> suspensions under UV irradiation [50]. Deltamethrin degradation increased in the presence of catalytic Cu [42]. The same trend was also observed by Burrows et al. [51], who evaluated the degradation ratio of the pesticide malathion by applying natural solar illumination. The 2% WO3/TiO<sup>2</sup> photocatalyst displayed the best photocatalytic efficiency [52]. Phosalone photodegradation effectiveness was influenced by irradiation time and the amount of TiO<sup>2</sup> present [43]. According to Liu et al. [53], the TiO2/HZSM-11 (30%) catalyst was effective in solution; it maintained its photocatalytic ability after many cycles, and it could be removed easily from the treated solution and reused immediately, giving it a great advantage for photocatalytic wastewater treatment. A recent study noted that the breakdown of the pesticides profenofos and triazophos was enhanced by TiO2/Ce application on the leaves of *Brassica chinensis* [54]. Nguyen and Juang [55] noted that TiO<sup>2</sup> use increased UV efficiency in p-chlorophenol degradation. Additionally, such a catalyst might be efficient under solar radiation, conserving electrical energy and consequently becoming an option for environmental remediation. Degradation rates were different in all studied conditions since wavelengths, exposure time, and solvent systems might affect the photodegradation [56]. It is worth noting the mechanism of photodegradation with and without a catalyst, since variations were noted between both conditions. In the remediation process without a photocatalyst, the pesticide molecules become excited by absorption of light energy of the UV radiation, causing homolysis, heterolysis, or photoionization. Whereas, in the process with a photocatalyst, the UV light energy will be absorbed by a semiconductor catalyst (titanium dioxide) to be photoexcited. However, a photoexcitation of the semiconductor catalyst occurs when the adsorbed light energy is greater than or at least equal to that of the gap between conduction and valence bands in the catalyst, leading to electron excitation to the conduction band (e−) and a positive hole (h+) in the valance band. Thus, oxidation–reduction reactions of the pesticide can be started by the radiation on the surface of semiconductor photocatalyst [52]. On the other hand, hydrogen peroxide (H2O2) as a powerful oxidant can be added to TiO<sup>2</sup> catalyst to enhance the effectiveness of the treatment by generating electrons, which leads to avoid the recombination of (e−)–(h+) pairs formed in the photocatalytic remediation [57]. This addition could reduce the effectiveness of the degradation process by modifying the photocatalyst surface by H2O<sup>2</sup> adsorption [58] and the inhibition of generated (h+) and reaction with hydroxyl radicals [59].

Additionally, studies on photoelectrochemical and catalysts applied for advanced treatment of wastewater are still at early stages despite the growing scientific and practical interest in this technology. Several recent studies have demonstrated that advanced oxidation processes (AOP) are more efficient for wastewater treatment, such as electrocatalysis, electro-fenton or photocatalysis, because hydroxyl radicals (OH•) are strong oxidizing agents that are generated from AOP under mild conditions.

Thus, the AOP have recently attracted the attention of researchers because they allow for the continuous electrocatalytic generation of strong oxidizing species under mild conditions. Moreover, energy can be saved by using sunlight in photovoltaic electrolysis systems and using a catalyst to speed up reactions [60]. Generally, UV is a well-known disinfection process normally used for drinking water treatment via their breakdown of water H-O bond. Consequently, water breakdown provides the strong oxidant HO• that has high potential as a redox and organic pollutant oxidizer [61,62]; therefore, UV are efficient in pesticides' removal or reduction from water. Interestingly, although the mode of action for UV in photolysis could be the same in relation to both tested pesticides in the current study, pesticides in varied media with different organic components responded differently. The type of dissolved organic materials in water may affect the UV absorption and it is expected that higher organic compounds in water lead to a high ability in UV absorbance and, therefore, high degradation ability is expected. Ag.WW was approved as a good medium for pesticide removal when the catalyst was added, and such a finding could be explained by the fact that Ag.WW had high organic constituents that could be good substrates for pesticide residues' conjugation. Since organic molecules have a high tendency towards UV absorbance, high degradation ability is therefore expected for the organic-pesticides' conjugate [63].

#### **4. Conclusions**

As consequences of industrial wastes and agricultural usage, pesticide residues in water are considered as one of the most serious environmental problems worldwide and, therefore, an efficient method for their elimination is needed. This study demonstrated the efficiency of photocatalytic agents for analyzing pesticide residues, and this is the first study on pesticide degradation (dieldrin and deltamethrin) using UV in three different water media collected from Saudi Arabia. UV radiation was used at 254 and 306 nm to induce photodegradation with and without photocatalytic TiO2. The results showed that UV use led to successful pesticide photolysis. For both tested pesticides, UV at 306 nm increased photolysis in a time-dependent manner. The catalyst increased the efficiency of UV irradiation at both wavelengths. The photolysis conditions were effective for both insecticides. Deltamethrin showed a higher degradation than dieldrin under all studied conditions. The obtained results in this study are very encouraging, so further kinetics studies of photo-remediation are recommended.

**Author Contributions:** Conceptualization, M.H.E.-S. and M.A.; Data curation, K.A., A.S.A. and M.A.; Formal analysis, M.O.A. and M.A.; Methodology, M.H.E.-S. and M.A.; Project administration, M.H.E.- S.; Resources, M.O.A., K.A. and A.S.A.; Supervision, M.H.E.-S., M.O.A. and M.A.; Writing—original draft, M.O.A., K.A., A.S.A. and M.A.-A.; Writing—review and editing, M.H.E.-S., M.O.A. and M.A. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was supported by the Chair of Environmental Pollution Research at Princess Nourah bint Abdulrahman University (Grant no. EPR023).

**Data Availability Statement:** All data supporting our findings are contained within the manuscript. Further details can be provided upon written request to the corresponding author.

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

#### **References**


**Ilil Levakov <sup>1</sup> , Yuval Shahar 1,2 and Giora Rytwo 1,2,\***


**Abstract:** Carbamazepine (CBZ) is one of the most common emerging contaminants released to the aquatic environment through domestic and pharmaceutical wastewater. Due to its high persistence through conventional degradation treatments, CBZ is considered a typical indicator for anthropogenic activities. This study tested the removal of CBZ through two different clay-based purification techniques: adsorption of relatively large concentrations (20–500 µmol L−<sup>1</sup> ) and photocatalysis of lower concentrations (<20 µmol L−<sup>1</sup> ). The sorption mechanism was examined by FTIR measurements, exchangeable cations released, and colloidal charge of the adsorbing clay materials. Photocatalysis was performed in batch experiments under various conditions. Despite the neutral charge of carbamazepine, the highest adsorption was observed on negatively charged montmorillonite-based clays. Desorption tests indicate that adsorbed CBZ is not released by washing. The adsorption/desorption processes were confirmed by ATR-FTIR analysis of the clay-CBZ particles. A combination of synthetic montmorillonite or hectorite with low H2O<sup>2</sup> concentrations under UVC irradiation exhibits efficient homo-heterogeneous photodegradation at µM CBZ levels. The two techniques presented in this study suggest solutions for both industrial and municipal wastewater, possibly enabling water reuse.

**Keywords:** carbamazepine; adsorption; clay minerals; organoclays; advanced oxidation processes; photocatalysis; water reuse

#### **1. Introduction**

Carbamazepine (CBZ) is one of the most common emerging contaminants released into the aquatic environment through domestic and pharmaceutical wastewater [1]. It is mainly used for epilepsy and bipolar disorder treatments and is considered a typical indicator for anthropogenic activities due to its high persistence through degradation processes in regular wastewater treatments. The removal of carbamazepine during conventional wastewater treatment processes was found to be neglectable and didn't exceed 10% [1–3]. Therefore, carbamazepine is found worldwide in surface water, groundwater, soil, and even drinking water with various concentrations of up to 10 µg L−<sup>1</sup> [4–7], and is expected to be found at higher levels in industrial wastewaters related to its manufacture and use (pharmaceutical and hospitals wastewater). Although no significant health risks were found associated with the exposure to carbamazepine residues in drinking water, several studies examined the negative side effect of consuming carbamazepine medicinally during pregnancy [8–10]. In addition, the ecotoxicity of carbamazepine for different aquatic species was demonstrated in many studies, revealing potential risks such as an increase in mortality rate, inhibition of growth, reproduction, and mobility [11–15]. The official regulations regarding carbamazepine in drinking water are limited and not available in most countries [7,16,17].

Over the years, various treatment approaches were tested aiming for efficient removal of carbamazepine from natural water bodies and wastewater. As mentioned above, removal

**Citation:** Levakov, I.; Shahar, Y.; Rytwo, G. Carbamazepine Removal by Clay-Based Materials Using Adsorption and Photodegradation. *Water* **2022**, *14*, 2047. https:// doi.org/10.3390/w14132047

Academic Editors: Martin Wagner and Sonja Bauer

Received: 26 May 2022 Accepted: 24 June 2022 Published: 26 June 2022

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

**Copyright:** © 2022 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 (https:// creativecommons.org/licenses/by/ 4.0/).

by conventional technologies was found to be neglectable, but few advanced approaches were able to remove CBZ with relatively high efficiency. For example, integrating biological modification with activated sludge increased the removal rate [18–20], specific microorganisms were found more efficient for its degradation [21,22], and enzymatic degradation including immobilization of the enzymes for increasing their operational stability [23]. In addition, advanced physicochemical treatment technologies such as nanofiltration (NF) and reverse osmosis (RO) were found to be effective for CBZ removal [24,25] even though the treatment of the concentrated brine afterward should be considered, and studies on that were not reported. Advanced oxidation processes [26,27] were an efficient option. Despite the high efficiency of those approaches, several problems limit their applicability, such as high costs, the toxicity of by-products in the oxidation process, biofouling, and the negative influence of natural organic matter during the removal by RO and NF [28–31].

Adsorption is one of the widest-used approaches for CBZ removal. Activated carbon is one of the common methods for adsorbing organic pollutants in drinking water, including pharmaceutically active compounds. The removal of carbamazepine by carbonbased sorbents was tested at various conditions with relatively high sorption capacities of up to 2 mmol g−<sup>1</sup> [32–34]. Different clays and organoclay are also used as potential adsorbents for pharmaceutical pollution as a low-cost and effective technique. Adsorption of carbamazepine on various clays was examined in several studies. Adsorption studies on montmorillonite are inconclusive, while some studies present poor-to-low adsorption capacities (0–0.02 mmol g−<sup>1</sup> ) [35–38] and others report considerably higher values (up to 0.15 mmol g−<sup>1</sup> ) [39,40]. Most studies have ascribed adsorption of CBZ on montmorillonite to Van der Waals interactions between the aromatic rings and the clay surface, and hydrogen bonds coordinating between oxygen atoms and exchangeable cations [36,38,40]. Khazri et al. (2017) have demonstrated S-type isotherms in pharmaceuticals adsorption [40], meaning low affinity at low concentration but following initially adsorbed molecules, promotes subsequently increased adsorption by Van der Waals forces between the pollutant aromatic moieties themselves. Some studies indicate that the adsorption occurred only on the clay's external surface and CBZ did not enter the clays' interlayer space [36].

Photocatalysis, an advanced oxidation process (AOP), is an additional approach for removing carbamazepine from contaminated water bodies. While photolysis techniques without a catalyst provided relatively poor removal performances, the addition of a catalyst was found to improve the degradation rate significantly [7,31,41]. The most common heterogeneous photocatalysts for CBZ photodegradation are catalytic grade oxides such as TiO2, ZnO, and MoS<sup>2</sup> [27,42,43]. The main disadvantages of using such conventional catalysts are the difficulties with separation of the particles and their reuse, weak stability, and low quantum efficiency [44–46]. Therefore, modification and improvement of the bare catalysts are of high interest in water treatment research. In recent years, several studies have demonstrated the applicability of using various clay-based materials as improved photocatalysts that provide efficient and stable reactions [44,47–50]. Furthermore, irradiation of UVC light with a combination of a homogeneous catalyst as hydrogen-peroxide and clay-based heterogeneous catalysts may deliver an effective advanced oxidation process, as was recently shown for BPS [51]. Thus, low price, high adsorption capability, stable structure for regeneration, and distinctive spatial structure for possible modification may turn clay-based materials into optimal potential efficient photocatalysts.

This study reports the removal of carbamazepine through two different clay-based purification techniques, aiming for different purposes: adsorption of relatively large concentrations (up to 500 µmol L−<sup>1</sup> ), focusing on industrial effluents and brine from filtration devices and photocatalysis of lower concentrations (<20 µmol L−<sup>1</sup> ), aiming for complete removal in domestic wastewater. The first part of the article includes CBZ adsorption isotherms on natural or modified clay minerals (several smectites, sepiolite, and hydrophobically modified montmorillonite). Fit to Langmuir, dual-mode and Sips models [52,53] was evaluated, and interpretation of the sorption mechanism is presented based on measuring the cations released during the processes, FTIR, and colloidal particle charge of CBZ-clay. The second part includes photocatalysis of carbamazepine using UVC irradiation and combinations of homo- and heterogeneous synthetic clay catalysts. Comparison with high quality catalytic grade TiO2, which is considered a "gold standard" of heterogeneous catalysts [54–56] was performed. The main objective of the research was to suggest a solution for carbamazepine removal, a persistent emerging contaminant, in an effective, low-cost, and reliable manner, at both mM and µM concentrations, that might be applied to reuse both industrial and municipal wastewater.

#### **2. Materials and Methods**

#### *2.1. Adsorption Experiments*

#### 2.1.1. Materials

Clay minerals used for the adsorption study were S9 "Pangel" sepiolite purchased from Tolsa SA (Madrid, Spain), bentonite (commercial montmorillonite, CAS: 1302-78-9) from Sigma-Aldrich (Rehovot, Israel), and Ca-montmorillonite prepared from SWy-1 clay (purchased from the Source Clays Repository of The Clay Minerals Society, Chantilly, VA, USA) using a batch procedure [57]. Thiamine hydrochloride (B1, CAS: 67-03-8), benzalkonium (bzk) solution (50% in H2O, CAS: 63449-41-2), and carbamazepine (CAS: 298-46-4) were supplied by Sigma-Aldrich (Israel).

#### 2.1.2. Clay Minerals Modification and Organoclay Preparation

All clay and organoclay matrices were prepared according to the same procedure with a concentration of 1% (10 g clay L−<sup>1</sup> ). For the clay suspensions, 1 g of the relevant clay was gradually added to 100 mL of double-distilled water (DDW) while stirring with magnetic stirring until homogenous suspensions were obtained. For the organoclay suspensions, different organic cations were added to the homogeneous clay suspensions and the complex was agitated for 24 h in order to reach equilibrium. The bentonite-B1 organoclay (bent-B1) was prepared by the addition of 175 g of thiamine powder to the bentonite suspension for a final load of 0.66 mmol thiamine g−<sup>1</sup> of clay. The bentonite-bzk organoclay (bentbzk) was prepared by adding 5.358 mL of benzalkonium solution (50%) for a total load of 0.6 mmol g−<sup>1</sup> . Previous studies showed that at those loads there is no release of the adsorbed organocations. Adsorption of the organic modifier was estimated by mass balance, after measuring concentrations in the liquid phase. To ensure complete removal of non-strongly bound modifiers from the complexes, the suspensions were centrifuged (3000 rpm for 30 min) and 90% of the supernatant was replaced with DDW. This procedure was repeated three times. Concentrations of the modifiers in supernatants were below detection limits.

#### 2.1.3. Adsorption Isotherms

Several batch experiments were conducted to study the adsorption of carbamazepine on various clay-based matrices. Each experiment included a set of different carbamazepine concentrations added to the relevant clay or organoclay with three replicates for each concentration. The experiments were conducted in 50-mL plastic tubes with a constant amount of adsorbent (0.25–1 g L−<sup>1</sup> , Table 1), added concentrations of carbamazepine ranging from 0 to 0.45 mM (0–106 mg L−<sup>1</sup> ), and DDW added to a final volume of 50 mL. Details on concentrations of carbamazepine and the relevant matrices in each experiment are described in Table 1. After the addition of carbamazepine, the samples were kept at room temperature (23 ± 1 ◦C) on an orbital shaker (100 rpm) for 24 h to reach equilibrium. The equilibrium was confirmed by additional sampling and analysis after 48 h. To separate the solids from the supernatant, 2 mL from each tube were sampled to Eppendorf vials and centrifuged at 15,000 rpm for 25 min in a SciLogex D2012 Eppendorfs (Rocky Hill, CT, USA) centrifuge. The concentration of carbamazepine and thiamine (in bent-B1 clay) were measured in the liquid phase using a diode-array HP 8452A UV–Vis spectrophotometer (Hewlett-Packard Company, Palo Alto, CA, USA) and determined by the absorbance

(OD = optical density) at 286 nm and 237 nm, respectively. As a preliminary experiment, CBZ spectrum was measured under different pH values and found stable through the range of 1.3–13 (results not shown). The adsorbed amount of CBZ was estimated by mass balance, thus subtracting the remaining concentration in the supernatant from the initial addition of carbamazepine. Average and standard deviation values were calculated from the triplicates, for the measured concentration in equilibrium (*X*-axis in the isotherm) and the adsorbed CBZ (*Y*-axis in the isotherm).


**Table 1.** Detailed adsorption experiments' conditions.

#### 2.1.4. FTIR Analysis

All sample suspensions were lyophilized (Christ Alpha 1-2 LD Plus, Germany) and the solids were analyzed in an attenuated total reflectance Fourier transforms infrared (ATR-FTIR) spectrophotometer. Analysis was performed on a Nicolet iS10 FTIR (Nicolet Analytical Instruments, Madison, WI, USA), using a SMART ATR device with a diamond crystal plate (Thermo Fisher Scientific, Madison, WI, USA) within a range of wavenumbers of 4000 to 500 cm−<sup>1</sup> . Spectra were recorded at 4 cm−<sup>1</sup> nominal resolution with mathematical corrections yielding a 1.0 cm−<sup>1</sup> actual resolution and averaged value from 50 measurements. The absorption intensity at different wavenumbers along the spectrum was quantified using TQ analyst EZ 8.0.2.97 software (Thermo Fisher Scientific). Quantification of adsorbed carbamazepine was based on the ratio between the absorbance intensities of specific absorption bands describing the sorbent in the case and CBZ [58]. In order to assess the stability of the CBZ adsorption, a release test was conducted subsequently to the adsorption experiments. The test was performed on three concentrations along the adsorption isotherms of CBZ to bentonite, Ca-SWy1, and bent-B1. For desorption experiments, CBZ loaded samples were washed three times, followed by centrifugation (3000 rpm for 30 min) and replacement of 90% of the supernatant with DDW. Washed samples were lyophilized and analyzed again in the ATR-FTIR spectrophotometer.

#### 2.1.5. Particle Charge Density Measurements

The colloidal charge of the particles was measured using a particle charge detector (PCD) (BTG Mütek, PCD-05, Eclépens, Switzerland). The PCD was connected to an automatic titration unit (PCD-05 Travel Titrator) with polyelectrolytes. The particle charge measurement is based on electrodes measuring the colloidal charge of a suspension agitated mechanically by a piston, in combination with titration of a charge-compensating polyelectrolyte [59,60]. Poly-DADMAC (poly-diallyl-dimethyl-ammonium chloride) or PES-NA (sodium-polyethylensulfonate) were used as cationic and anionic polyelectrolytes according to the charge of the particles. Each measurement required 10 mL of the homogeneous suspension and colloidal charge results were normalized according to the mass of clay or organoclay in each case.

#### 2.1.6. Exchangeable Cations Measurements

Examination of possible cations' exchange on the clay interlayers was evaluated by measuring the changes in cations concentrations in each of the supernatants of the adsorption points along the isotherm. Several major cations (calcium, sodium, potassium, and magnesium) were measured by inductively coupled plasma-optical emission spectrometry (ICP-OES) analysis. The analysis was performed with a Thermo Scientific IRIS Intrepid II XDL ICP-OES (Thermo Electron Corporation, Waltham, MA, USA). All samples were filtered (0.2 µm) and HNO<sup>3</sup> was added to a final concentration of 2%. Multielement standard solution (multi-3 for ICP, 49596 Sigma-Aldrich) was used for calcium, magnesium, potassium, and sodium calibration, which calibration curves of 1–10, 0.2–2, 0.1–1, and 0.5–5 mg L−<sup>1</sup> respectively.

#### 2.1.7. Adsorption Models

As a first approximation, the fit of all adsorption isotherms to the Langmuir adsorption equation (Equation (1))

$$\text{Cs} = \frac{\text{S}\_{\text{max}} \text{K}\_L \text{C}\_L}{1 + \text{K}\_L \text{C}\_L} \tag{1}$$

was tested. *Cs* is the sorbed concentration (mmol g−<sup>1</sup> ), *C<sup>L</sup>* is the solution concentration (mM), *S*max defines the number of adsorption sites per mass of sorbent (mmol g−<sup>1</sup> ), and *K<sup>L</sup>* is the Langmuir adsorption coefficient (mM−<sup>1</sup> ). We also tested fit to the Dual-mode model (DMM) (Equation (2)), which combines the Langmuir equation and a linear equation to simulate the combination of a site-specific adsorption mechanism and partitioning mechanism occurring simultaneously [53,61]. Nevertheless, the linear equation component was found to be neglectable, therefore only results of the Langmuir model are reported.

Since some of the sorbents exhibit type V sigmoidal (S) adsorption isotherms, we tested also suitable models for such behavior as BET, Klotz, and Sips equations [62]. From those models we chose Sips equations since it was the only model that showed improved fit.

Sips model is a hybrid combination of Langmuir and Freundlich equations that can describe Type V behavior [63]. Sips model can be described by Equation (2).

$$\text{Cs} = \frac{\text{S}\_{\text{max}} (K\_L \text{C}\_L)^{B\_s}}{1 + (K\_L \text{C}\_L)^{B\_s}} \tag{2}$$

in which *B<sup>s</sup>* is known as "Sips model exponent" [64]. The equation appears in the literature with different notations, and in some cases *K<sup>L</sup>* is not included in the exponent [63]. We adopted the notation in Equation (2) [62] since it keeps the units of *K<sup>L</sup>* identical to Langmuir equation.

The non-linear curve fitting was performed using scipy.optimize.curve\_fit functions from the SciPy package (version 1.4.1), Python (Python 3.7.13), https://scipy.org/citingscipy/ (accessed on 20 June 2022). The function calculating the specific model parameters (*S*max, *K<sup>L</sup>* and *BS*) for each isotherm.

#### *2.2. Photocatalysis Experiments*

#### 2.2.1. Materials

Catalyst-grade industrial high-quality TiO<sup>2</sup> (Hombikat®, American Elements, Los Angeles, CA, USA) and a 30% (9.79 M) concentrated H2O<sup>2</sup> solution were purchased from Merck\Sigma-Aldrich (Merck KGaA, Darmstadt, Germany). SYn-1 Barasym SSM-100 synthetic mica-montmorillonite was obtained from the Source Clays Repository of The Clay Minerals Society (Chantilly, VA, USA), whereas Laponite-RD was provided by BYK-Chemie GmbH (Wesel, Germany).

#### 2.2.2. Methods

Degradation of CBZ from a 21.2 µM (5 mg L−<sup>1</sup> ) solution was measured in batch experiments in 100-mL UV-C-transparent quartz glass (refractive Index n = 1.5048), 5.3-cm diameter beaker placed in a Rayonet RMR-600 mini photochemical chamber reactor (Southern New England Ultraviolet Company, Branford, CT, USA), as described in previous studies [65,66]. The photoreactor was equipped with eight RMR 2537A lamps (254 nm wavelength), each lamp emitting an average irradiance flux of 19 W m−<sup>2</sup> at 254 nm, equivalent to an overall intensity of 152 W m−<sup>2</sup> , as measured in the center of the chamber using a Black Comet SR spectrometer with an F400 UV–VIS–SR-calibrated fiber optic probe equipped with a CR2 cosine light receptor (StellarNet Inc., Tampa, FL, USA). The Black Comet SR spectrometer was also used to measure the spectrum of the solutions during experiments using a 20 mm pathlength DP400 dip probe cuvette (StellarNet Inc., Tampa, FL, USA) placed inside the beaker. The solutions were constantly mixed with an external stirrer (VELP Scientifica, Usmate Velate, Italy) rotating at 100 rpm. Spectra were measured using the SpectraWiz software (StellarNet Inc., Tampa, FL, USA) every 10–20 s for approximately 20–60 min (depending on the experiment). A short MP4 clip showing the experimental setup is available as Supplementary Material. The measurement procedure led to >150–400 data points for each experiment. Data were transformed to comma-separated values (CSV) files, and absorption of the net absorption of CBZ at 286 nm (ε<sup>286</sup> = 12,826 M−<sup>1</sup> cm−<sup>1</sup> ) was evaluated and downloaded after correcting the baseline. To allow comparison between parameters in different reaction mechanisms, the "relative dimensionless concentration at time t" [A(t)] was evaluated [67] as Ct/C<sup>0</sup> = ODt/OD<sup>0</sup> (the ratio of actual to initial concentration, or actual to initial light absorbance); thus A<sup>0</sup> = 1. Analysis of the data was performed as described in Section 2.2.3.

HPLC Chromatography measurements were kindly performed as outsourcing by Dr. Sara Azarred at Shamir Research Institute, to confirm that quantification using UV-Visible measurements yields indeed reliable results. HPLC measurements indeed confirmed (results not shown) that direct UV-Visible spectroscopy measurements are accurate and effective at the required concentrations range (0.1–5 mg L−<sup>1</sup> , ~0.5–25 µM) and presence of by-products does not influence the measurements.

The experiments performed included homogenous photocatalysis of CBZ with different concentrations (0.5–2.5 mg L−<sup>1</sup> , 14.7–73.5 µM) of H2O2, heterogeneous photocatalysis of CBZ with various concentrations (0.2–1 mg L−<sup>1</sup> ) of TiO2, barasym or laponite, and combined hetero-homogeneous photocatalysis of CBZ with 0.5 or 2.0 mg L−<sup>1</sup> H2O<sup>2</sup> and heterogeneous catalysts (TiO2, barasym, laponite) at 0.2 mg L−<sup>1</sup> .

#### 2.2.3. Analysis of the Data

In order to compare the efficacy of the different photocatalysis processes, an evaluation of the pseudo order, the kinetic rate, and the half-life of each process was performed. Calculations were done using the procedure extensively described in previous studies [51,68]. Considering the rate of change in concentration follows Equation (3):

$$\upsilon = \frac{d[A]}{dt} = -k\_a[A]^{n\_a} \tag{3}$$

where *υ* is the reaction rate, *k<sup>a</sup>* is the apparent rate coefficient, *A* is the concentration of the pollutant in case, and *n<sup>a</sup>* is the apparent or "pseudo" reaction order [69], the concentration at time *t* can be calculated if the kinetic rate coefficient *k<sup>a</sup>* and the pseudo-order *n<sup>a</sup>* are known(as long as *n<sup>a</sup>* 6= 1), using:

$$[A]\_{(t)} = \left(\frac{1}{\frac{1}{[A\_0]^{n\_a-1}} + (n\_a-1)k\_at}\right)^{\frac{1}{n\_a-1}}\tag{4}$$

"Half-life time" (*t*1/2), defined by the time it takes for the concentration of a reactant to reach half of its initial value [69], are easy-to-compare parameters, even in cases where pseudo orders are completely different. Half-life times can be evaluated by solving mathematically Equations (3) and (4) to the case were [*A*](*t*) = 0.5, yielding for *n<sup>a</sup>* 6= 1

$$t\_{\frac{1}{2}} = \frac{2^{n\_d - 1} - 1}{(n\_d - 1)k\_d[A\_0]^{n\_d - 1}}\tag{5}$$

It should be emphasized that except for "first-order" processes, half-life times strongly depends on the initial concentration, as seen in Equation (5). This should be considered when comparing the efficiency of processes, and the use of a constant initial concentration of pollutants is important.

Pseudo-orders and the kinetic coefficient that exhibits the best fit to each of the treatments were found as described in previous studies by a "bootstrap" [70,71] procedure based on choosing five random sets of 20 values from the several hundreds of data points in each experiment, and fitting the optimal parameters using the Solver tool in Excel® software [68].

#### **3. Results and Discussion**

#### *3.1. Adsorption Isotherms*

Adsorption isotherms of carbamazepine on the different sorbents are presented in Figure 1A. The adsorption was tested as described in Section 2.1.3 on five clay-based adsorbents, including three raw clays: (bentonite, Ca-SWy1, and sepiolite), and two organoclays prepared as described in Section 2.1.2 (bent-bzk and bent-B1). The adsorption isotherms were also described by the Langmuir adsorption model, and the parameters *S*max and *K<sup>L</sup>* were evaluated (Table 2). R<sup>2</sup> and RMSE values between the observed and the calculated adsorption results indicate a good fit with the Langmuir model (Table 2, R<sup>2</sup> > 0.99). It is important to mention that the following Langmuir models are relevant mainly for the high CBZ adsorption results (Figure 1A). While the adsorption isotherm on bentonite and Ca-SWy1 at low CBZ concentration (Figure 1B) presented a Type V isotherm behavior, that did not fully fit the standard adsorption models and required a more advanced model to describe the slight S-shape measured. For that purpose we tested several models suitable for Type V isotherms as BET, Sips, and Klotz [62]). The only relative improvement in the fit was obtained by the Sips model that is described in Section 2.1.7. It should be emphasized that as shown in Table 2, the improvement in the fit when compared with Langmuir model is minimal, and in the case of the organoclays (BZK- and B1 bentonites) there is no improvement at all, and the exponent in the Sips model (*Bs*) for those sorbents is close to 1.

Adsorption of CBZ on sepiolite is negligible (Figure 1A). This is not obvious, since several non-charged molecules and oils are adsorbed in large amounts on this clay [72,73]. As for the smectites, organoclay based on BZK exhibited the lowest adsorption, whereas raw bentonite and Ca-SWy-1 adsorb more than 0.5 mmol g−<sup>1</sup> . Such effect is unusual since smectites usually have a low affinity to non-charged organic molecules. However, it should be emphasized that at low CBZ concentration adsorption to those clays is almost zero (see Figure 1B, which shows adsorption isotherms at low CBZ concentrations), yielding an "S-type" (or "Type V") isotherm [74]. This indicates that the direct interaction between the clay surface and the pollutant is low, and only after obtaining some coverage of the clay surface, adsorption increase. A similar observation was made for several pharmaceuticals in previous studies [40]. On the other hand, bent-B1 maximum adsorption is lower (app. 0.25 mmol g−<sup>1</sup> , see Figure 1A) but it is also very efficient at low concentrations (see Figure 1B). The affinity of carbamazepine to bent-B1 increased considerably as demonstrated also in a higher calculated Langmuir *K<sup>L</sup>* value, 16.9 compared to 2.89 mM−<sup>1</sup> on raw bentonite (Table 2). Due to the neutral charge of carbamazepine, it is assumed that more effective adsorption will be observed on the neutral charge particles such as bent-B1 and bent-bzk organoclays. However, higher adsorption capacities were found

g−1 [75–78].

ganoclays.

organoclay

for montmorillonite clays. Studies reporting adsorption of CBZ on clay minerals mention lower adsorption values as 0.02–0.2 mmol g−<sup>1</sup> [35,36,39]. Few studies have evaluated the adsorption on modified and organo-clays with even lower adsorbed amounts, from 0–0.05 mmol g−<sup>1</sup> [75–78]. known spectrums of the components (B1 and carbamazepine) at known concentrations, and calculating the mix spectrum by superposition [80] using the "Solver"® optimization tool of the Microsoft Excel computer program.

(see Figure 1B). The affinity of carbamazepine to bent-B1 increased considerably as demonstrated also in a higher calculated Langmuir *KL* value, 16.9 compared to 2.89 mM−<sup>1</sup> on raw bentonite (Table 2). Due to the neutral charge of carbamazepine, it is assumed that more effective adsorption will be observed on the neutral charge particles such as bent-B1 and bent-bzk organoclays. However, higher adsorption capacities were found for montmorillonite clays. Studies reporting adsorption of CBZ on clay minerals mention lower adsorption values as 0.02–0.2 mmol g−1 [35,36,39]. Few studies have evaluated the adsorption on modified and organo-clays with even lower adsorbed amounts, from 0–0.05 mmol

Despite bent-B1 advantage in adsorption at low concentration, some thiamine (B1) was released from the bent-B1 complex during the adsorption process. The concentration of the released B1 was very low (averagely 0.02 mM) with relatively constant values through different CBZ adsorption concentrations. Moreover, thiamine is considered a non-hazardous component that can provide a safe use under high standard regulations

eration. B1 release interfered with UV Visible measurements, and in order to overcome this problem a simple mathematical spectra separation was performed, based on two well-

*Water* **2022**, *14*, x FOR PEER REVIEW 18 of 28

**Figure 1.** Adsorption isotherms of (**A**) carbamazepine on bentonite (blue asterisk), sepiolite (red triangle), Ca-SWy1 (black circle), bent-bzk (green square), and bent-B1 (purple diamond). Error bars represent triplicate standard deviations. The dashed lines represent Langmuir model predictions according to the estimated parameters detailed in Table 2. (**B**) Isotherms at low concentration of carbamazepine on bentonite, Ca-SWy1, and bent-B1. **Figure 1.** Adsorption isotherms of (**A**) carbamazepine on bentonite (blue asterisk), sepiolite (red triangle), Ca-SWy1 (black circle), bent-bzk (green square), and bent-B1 (purple diamond). Error bars represent triplicate standard deviations. The dashed lines represent Langmuir model predictions according to the estimated parameters detailed in Table 2. (**B**) Isotherms at low concentration of carbamazepine on bentonite, Ca-SWy1, and bent-B1.

**Table 2.** Langmuir and Sips equations' parameters for carbamazepine adsorption on clays and or-**Table 2.** Langmuir and Sips equations' parameters for carbamazepine adsorption on clays and organoclays.


Despite bent-B1 advantage in adsorption at low concentration, some thiamine (B1) was released from the bent-B1 complex during the adsorption process. The concentration of the released B1 was very low (averagely 0.02 mM) with relatively constant values through different CBZ adsorption concentrations. Moreover, thiamine is considered a non-hazardous component that can provide a safe use under high standard regulations [79]. However, such instability in the behavior of the sorbent should be taken into consideration. B1 release interfered with UV Visible measurements, and in order to overcome this problem a simple mathematical spectra separation was performed, based on two well-known spectrums of the components (B1 and carbamazepine) at known concentrations, and calculating the mix spectrum by superposition [80] using the "Solver"® optimization tool of the Microsoft Excel computer program.

Sips 0.318 18.20 1.05 0.998 0.004

#### *3.2. Colloidal Charge 3.2. Colloidal Charge*

The influence of CBZ adsorption on the electrokinetic colloidal charge of bentonite, Ca-SWy1, and bent-B1 organoclay was evaluated (Figure 2) using a particle charge detector (PCD) as described in Section 2.1.5. The initial colloidal charge of bentonite was significantly more negative than Ca-SWy1 due to the influence of the divalent calcium ions increasing neutralization on the Stern layer. Thus, in raw bentonite, the presence of monovalent sodium ions resulted in a more negative electrokinetic charge of the clay particles. The colloidal charge of the organoclay (bent-B1) was around zero due to the exchange of B1 with the inorganic ions, forming a non-charged surface. Similar values were observed in organoclays with organocations added at amounts near the cation exchange capacity of the clay as in B1- [52], berberine- [81], crystal violet, and tetraphenyl-phosphonium [82] smectites. For the raw bentonite and Ca-SWy-1, the increase in the electrokinetic colloidal charge, making it closer to neutralization, is attributed to the adsorption of the hydrophobic carbamazepine molecule. In raw bentonite, the increase in charge is accompanied by a cationic exchange of sodium with additional calcium arriving from the carbamazepine solution. Those results will be further discussed in the next chapters (Sections 3.3 and 3.4). The influence of CBZ adsorption on the electrokinetic colloidal charge of bentonite, Ca-SWy1, and bent-B1 organoclay was evaluated (Figure 2) using a particle charge detector (PCD) as described in Section 2.1.5. The initial colloidal charge of bentonite was significantly more negative than Ca-SWy1 due to the influence of the divalent calcium ions increasing neutralization on the Stern layer. Thus, in raw bentonite, the presence of monovalent sodium ions resulted in a more negative electrokinetic charge of the clay particles. The colloidal charge of the organoclay (bent-B1) was around zero due to the exchange of B1 with the inorganic ions, forming a non-charged surface. Similar values were observed in organoclays with organocations added at amounts near the cation exchange capacity of the clay as in B1- [52], berberine- [81], crystal violet, and tetraphenyl-phosphonium [82] smectites. For the raw bentonite and Ca-SWy-1, the increase in the electrokinetic colloidal charge, making it closer to neutralization, is attributed to the adsorption of the hydrophobic carbamazepine molecule. In raw bentonite, the increase in charge is accompanied by a cationic exchange of sodium with additional calcium arriving from the carbamazepine solution. Those results will be further discussed in the next chapters (Sections 3.3 and 3.4).

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**Figure 2.** Colloidal charge of bentonite (blue asterisk), Ca-SWy1 (black circle), and bent-B1 organoclay (purple diamond) at 0–1.5 mmol g<sup>−</sup>1 CBZ addition. **Figure 2.** Colloidal charge of bentonite (blue asterisk), Ca-SWy1 (black circle), and bent-B1 organoclay (purple diamond) at 0–1.5 mmol g−<sup>1</sup> CBZ addition.

#### *3.3. Cation Exchange in Raw Bentonite 3.3. Cation Exchange in Raw Bentonite*

The large adsorbed CBZ amounts on raw bentonite, which is, as most natural clay minerals, negatively charged, led to the assumption that CBZ may probably behave as a cation, exchanging other exchangeable cations from the raw clay. To test this assumption, ion exchange processes were examined as part of the adsorption mechanisms evaluation. Raw bentonite is reported to have a cation exchange capacity (CEC) of 0.8 mmole g−1, whereas the cations composition is similar to that measured in SWy-1 and SWy-2 clays, and thus about 30% of the CEC is Na+, while almost all the rest are divalent cations [57]. The release of inorganic cations due to the adsorption of CBZ was measured as described in Section 2.1.6. It is interesting to mention that apparently according to our measurements, carbamazepine as purchased from Sigma may contain traces of calcium in it. This assumption is based on Ca concentrations measured in "pure" CBZ solutions using ICP-OES analysis and reinforced by FTIR measurements reported in Section 3.4 that exhibit a strong peak ascribed to CaCO3. An additional explanation to the presence of such Ca traces in the CBZ stock could be related to a possible contamination during the laboratory work. In any case, Figure 3A represents the release of sodium and calcium ions into the liquid phase as a function of the adsorption of carbamazepine. The release was calculated by subtracting the ions in the initial clay solution from the various solution concentrations after carbamazepine adsorption. The release of exchangeable sodium cations from the bentonite is linearly correlated to the amount of adsorbed carbamazepine (Figure 3A). Despite this linear correlation, CBZ adsorption is not explained by the sodium release and The large adsorbed CBZ amounts on raw bentonite, which is, as most natural clayminerals, negatively charged, led to the assumption that CBZ may probably behave as a cation, exchanging other exchangeable cations from the raw clay. To test this assumption, ion exchange processes were examined as part of the adsorption mechanisms evaluation. Raw bentonite is reported to have a cation exchange capacity (CEC) of 0.8 mmole g−<sup>1</sup> , whereas the cations composition is similar to that measured in SWy-1 and SWy-2 clays, and thus about 30% of the CEC is Na<sup>+</sup> , while almost all the rest are divalent cations [57]. The release of inorganic cations due to the adsorption of CBZ was measured as described in Section 2.1.6. It is interesting to mention that apparently according to our measurements, carbamazepine as purchased from Sigma may contain traces of calcium in it. This assumption is based on Ca concentrations measured in "pure" CBZ solutions using ICP-OES analysis and reinforced by FTIR measurements reported in Section 3.4 that exhibit a strong peak ascribed to CaCO3. An additional explanation to the presence of such Ca traces in the CBZ stock could be related to a possible contamination during the laboratory work. In any case, Figure 3A represents the release of sodium and calcium ions into the liquid phase as a function of the adsorption of carbamazepine. The release was calculated by subtracting the ions in the initial clay solution from the various solution concentrations after carbamazepine adsorption. The release of exchangeable sodium cations from the bentonite is linearly correlated to the amount of adsorbed carbamazepine (Figure 3A). Despite this linear correlation, CBZ adsorption is not explained by the sodium release and

the ratio between carbamazepine adsorption and Na release was higher than 2. Thus, this is not a mere exchange CBZ/Na<sup>+</sup> . Moreover, since calcium was added to the solution through the carbamazepine stock solution (as mentioned above), a comparison between the calcium addition and the release of sodium was conducted (Figure 3B), resulted in a linear correlation with a 1:1 ratio as for mmolec. Thus, the hypothesis that CBZ exchanges Na<sup>+</sup> appears wrong, and a more reasonable explanation for the release of sodium cations is a Na+/Ca2+ exchange process (Ca2+ coming from the CBZ solution) considering there is a strong preference in clays for divalent cations [83], and the adsorption of carbamazepine did not include ions exchange on the clay's negative sites. To confirm this assumption, we performed the adsorption isotherms on Ca homoionic SWy-1 montmorillonite (Ca-SWy1), where all the CEC was *a-priori* saturated with Ca2+. As shown in Figure 1A CBZ adsorption to Ca-SWy1 reaches similar and even slightly larger values than for raw bentonite, with similar Langmuir and Sips *S*max values. the ratio between carbamazepine adsorption and Na release was higher than 2. Thus, this is not a mere exchange CBZ/Na+. Moreover, since calcium was added to the solution through the carbamazepine stock solution (as mentioned above), a comparison between the calcium addition and the release of sodium was conducted (Figure 3B), resulted in a linear correlation with a 1:1 ratio as for mmolec. Thus, the hypothesis that CBZ exchanges Na+ appears wrong, and a more reasonable explanation for the release of sodium cations is a Na+/Ca2+ exchange process (Ca2+ coming from the CBZ solution) considering there is a strong preference in clays for divalent cations [83], and the adsorption of carbamazepine did not include ions exchange on the clay's negative sites. To confirm this assumption, we performed the adsorption isotherms on Ca homoionic SWy-1 montmorillonite (Ca-SWy1), where all the CEC was *a-priori* saturated with Ca2+. As shown in Figure 1A CBZ adsorption to Ca-SWy1 reaches similar and even slightly larger values than for raw bentonite, with similar Langmuir and Sips *S*max values.

**Figure 3.** (**A**) Cations released to supernatant from bentonite (blue asterisk) or Ca-SWy1 (black circle) as a function of the adsorbed CBZ. (**B**) Na+ measured due to the addition of Ca+2 from CBZ solution (green diamonds). **Figure 3.** (**A**) Cations released to supernatant from bentonite (blue asterisk) or Ca-SWy1 (black circle) as a function of the adsorbed CBZ. (**B**) Na<sup>+</sup> measured due to the addition of Ca+2 from CBZ solution (green diamonds).

#### *3.4. FTIR Analysis 3.4. FTIR Analysis*

One of the hypotheses to explain the relatively large adsorbed amounts of neutral CBZ on negatively charged clays was that as the matter of fact it undergoes degradation on the surface of the mineral, as was observed for example in the case of tri-methyl aryl dyes "adsorbed" on Texas vermiculite (VTx-1) [84]. In order to test that, FTIR spectra of dried CBZ-clays were measured using an ATR device as described in Section 2.1.4. The rationale aimed to confirm and evaluate the adsorption of carbamazepine on the absorbent particles by identifying the relevant structural group in the measured samples, whereas CBZ degradation will lead to different functional group vibrations. Figure 4A shows the spectrum of CBZ, raw bentonite, and CBZ-bentonite at several carbamazepine amounts. CBZ-bentonite samples exhibit five absorption bands that were not observed in raw bentonite, at approximately 1640, 1570, 1490, and 1460–1435 cm−1. Those absorption bands were correlated to typical peaks of functional groups of CBZ carbamazepine, as observed in the carbamazepine spectrum sample, and are known from the literature [85]. The only absorption band in raw bentonite in this region is at ~1630 cm−1 and ascribed to O-H deformation of hygroscopic water [86]. While the three CBZ absorption bands in the range 1400–1500 cm−1 (ascribed to C=C vibrations) are in almost the same place for adsorbed and raw CBZ, bands ascribed to C-N bond (at app. 1600 cm−1) and to the amide group (NH2 scissoring/C=O stretching, at ~1670 cm−1) appear shifted to lower energies. This may indicate that the interaction between CBZ molecules and the clay is via the amide group. One of the hypotheses to explain the relatively large adsorbed amounts of neutral CBZ on negatively charged clays was that as the matter of fact it undergoes degradation on the surface of the mineral, as was observed for example in the case of tri-methyl aryl dyes "adsorbed" on Texas vermiculite (VTx-1) [84]. In order to test that, FTIR spectra of dried CBZ-clays were measured using an ATR device as described in Section 2.1.4. The rationale aimed to confirm and evaluate the adsorption of carbamazepine on the absorbent particles by identifying the relevant structural group in the measured samples, whereas CBZ degradation will lead to different functional group vibrations. Figure 4A shows the spectrum of CBZ, raw bentonite, and CBZ-bentonite at several carbamazepine amounts. CBZ-bentonite samples exhibit five absorption bands that were not observed in raw bentonite, at approximately 1640, 1570, 1490, and 1460–1435 cm−<sup>1</sup> . Those absorption bands were correlated to typical peaks of functional groups of CBZ carbamazepine, as observed in the carbamazepine spectrum sample, and are known from the literature [85]. The only absorption band in raw bentonite in this region is at ~1630 cm−<sup>1</sup> and ascribed to O-H deformation of hygroscopic water [86]. While the three CBZ absorption bands in the range 1400–1500 cm−<sup>1</sup> (ascribed to C=C vibrations) are in almost the same place for adsorbed and raw CBZ, bands ascribed to C-N bond (at app. 1600 cm−<sup>1</sup> ) and to the amide group (NH<sup>2</sup> scissoring/C=O stretching, at ~1670 cm−<sup>1</sup> ) appear shifted to lower energies. This may indicate that the interaction between CBZ molecules and the clay is via the amide group.

cess as the reason for CBZ decrease in the equilibrium solution.

firmed as stable processes without unexpected CBZ release.

**Figure 4.** (**A**) ATR-FTIR spectra of CBZ (black, OD values on the right axis), bentonite raw clay (blue), and bentonite with added CBZ (in mmol g<sup>−</sup>1) as denoted in the legend *(*OD values on the left axis). (**B**) Normalized absorption bands' height at 1490 and area at 1460–1435 cm<sup>−</sup>1 compared to the adsorption results as measured by the mass balance during the adsorption experiments. **Figure 4.** (**A**) ATR-FTIR spectra of CBZ (black, OD values on the right axis), bentonite raw clay (blue), and bentonite with added CBZ (in mmol g−<sup>1</sup> ) as denoted in the legend (OD values on the left axis). (**B**) Normalized absorption bands' height at 1490 and area at 1460–1435 cm−<sup>1</sup> compared to the adsorption results as measured by the mass balance during the adsorption experiments.

An increase in all absorption bands in the range 1400–1700 cm−1 is observed accordingly to the adsorption process as measured in the experiment (Figure 4A). Relative quantification of the CBZ absorbed can be performed according to Section 2.1.4, by calculating the ratio between the intensity or the area of CBZ absorption bands, to that of the structural O-H band related to the clay at 3620 cm−1, where CBZ has no absorption at all. Subsequently, those ratios were compared to the amount of carbamazepine adsorbed to the clay as measured in the previous adsorption experiment (Figure 1). Figure 4B represents the comparison of two normalized absorption bands to the adsorption results, including the absorption bands' height at 1490 and their area at 1460–1435 cm−1. The linear correlation with very good fit (R2 > 0.99) between the two data sets confirms the adsorption pro-

An additional absorption band was observed in the raw carbamazepine spectrum at approximately 1370 cm−1 and was not observed in any of the CBZ-bentonite samples (Figure 4A). According to the literature, these absorption bands may represent the presence of calcite (CaCO3) [86] in the carbamazepine or as a consequence of Ca impurities in our stock solution as described above. Presence of calcite is confirmed by a small and sharp absorption band at 870 cm−1 (results not shown) in the CBZ spectrum. The absence of this peak in all CBZ-bentonite spectra indicates that calcite was released to the liquid phase, as correlated to the increase in calcium ions that was observed after adsorption (Figure 3). The stability of the CBZ adsorption was examined by a release test (as described in Section 2.1.4). The results of the washed and original samples were compared, and no significant differences were observed in the absorption bands and the calculated ratios (results not shown). Hence, the adsorptions of CBZ to bentonite, Ca-SWy1, and bent-B1 were con-

*3.5. Photocatalytic Degradation of Carbamazepine*  3.5.1. Homogenous Photocatalysis with Increasing Concentrations of H2O2 Figure 5 shows the photodegradation of a 21.2 μM (5 mg L−1) CBZ solution, under UVC irradiation as described in Section 2.2.2, with 0.5–2.5 mg L−1 (14.7–73.5 μM) H2O2 as homogenous catalysts. Such H2O2 concentration is in the range that was used recently for photodegradation of caffeine [65], but considerably lower than was usually used for photo- [87] or radio-catalysis [88] of CBZ. It can be seen that CBZ is stable under UVC radiation and does not undergo any photolysis. At the initial CBZ concentration used in this study (21.2 μM) very low H2O2 concentration (0.5 mg L−1) exhibits almost no degradation, as in photolysis. Increasing the H2O2 concentrations to 1.0 mg L−1 changes the An increase in all absorption bands in the range 1400–1700 cm−<sup>1</sup> is observed accordingly to the adsorption process as measured in the experiment (Figure 4A). Relative quantification of the CBZ absorbed can be performed according to Section 2.1.4, by calculating the ratio between the intensity or the area of CBZ absorption bands, to that of the structural O-H band related to the clay at 3620 cm−<sup>1</sup> , where CBZ has no absorption at all. Subsequently, those ratios were compared to the amount of carbamazepine adsorbed to the clay as measured in the previous adsorption experiment (Figure 1). Figure 4B represents the comparison of two normalized absorption bands to the adsorption results, including the absorption bands' height at 1490 and their area at 1460–1435 cm−<sup>1</sup> . The linear correlation with very good fit (R<sup>2</sup> > 0.99) between the two data sets confirms the adsorption process as the reason for CBZ decrease in the equilibrium solution.

An additional absorption band was observed in the raw carbamazepine spectrum at approximately 1370 cm−<sup>1</sup> and was not observed in any of the CBZ-bentonite samples (Figure 4A). According to the literature, these absorption bands may represent the presence of calcite (CaCO3) [86] in the carbamazepine or as a consequence of Ca impurities in our stock solution as described above. Presence of calcite is confirmed by a small and sharp absorption band at 870 cm−<sup>1</sup> (results not shown) in the CBZ spectrum. The absence of this peak in all CBZ-bentonite spectra indicates that calcite was released to the liquid phase, as correlated to the increase in calcium ions that was observed after adsorption (Figure 3). The stability of the CBZ adsorption was examined by a release test (as described in Section 2.1.4). The results of the washed and original samples were compared, and no significant differences were observed in the absorption bands and the calculated ratios (results not shown). Hence, the adsorptions of CBZ to bentonite, Ca-SWy1, and bent-B1 were confirmed as stable processes without unexpected CBZ release.

#### *3.5. Photocatalytic Degradation of Carbamazepine*

#### 3.5.1. Homogenous Photocatalysis with Increasing Concentrations of H2O<sup>2</sup>

Figure 5 shows the photodegradation of a 21.2 µM (5 mg L−<sup>1</sup> ) CBZ solution, under UVC irradiation as described in Section 2.2.2, with 0.5–2.5 mg L−<sup>1</sup> (14.7–73.5 µM) H2O<sup>2</sup> as homogenous catalysts. Such H2O<sup>2</sup> concentration is in the range that was used recently for photodegradation of caffeine [65], but considerably lower than was usually used for photo- [87] or radio-catalysis [88] of CBZ. It can be seen that CBZ is stable under UVC radiation and does not undergo any photolysis. At the initial CBZ concentration used in this study (21.2 µM) very low H2O<sup>2</sup> concentration (0.5 mg L−<sup>1</sup> ) exhibits almost no degradation, as in photolysis. Increasing the H2O<sup>2</sup> concentrations to 1.0 mg L−<sup>1</sup> changes

the pseudo-order as evaluated using Equations (3) and (4) from n = 0 to n = 0.75, while *t*1/2 as evaluated using Equation (5) changes from 219 to 15.6 min. Further increase of H2O<sup>2</sup> to 2 or 2.5 mg L−<sup>1</sup> reduces *t*1/2 further to 7.6 and 6.4 min, respectively. Pseudo orders and half lifetimes of all experiments are shown also in Table 3. evaluated using Equation (5) changes from 219 to 15.6 min. Further increase of H2O2 to 2 or 2.5 mg L−1 reduces *t*1/2 further to 7.6 and 6.4 min, respectively. Pseudo orders and half lifetimes of all experiments are shown also in Table 3.

pseudo-order as evaluated using Equations (3) and (4) from n = 0 to n = 0.75, while *t*1/2 as

*Water* **2022**, *14*, x FOR PEER REVIEW 22 of 28

**Figure 5.** Photodegradation of 21.2 μM (5 mg L<sup>−</sup>1) carbamazepine solution under UVC irradiation at several concentrations of hydrogen peroxide. **Figure 5.** Photodegradation of 21.2 µM (5 mg L−<sup>1</sup> ) carbamazepine solution under UVC irradiation at several concentrations of hydrogen peroxide.


**Table 3.** Pseudo-orders and half-life times for photodegradation experiments. **Table 3.** Pseudo-orders and half-life times for photodegradation experiments.

0.2 0.5 1.04 ± 4.27% 37.0 ± 0.54% 3.5.2. Heterogenous Photocatalysis with TiO2, Barasym and Laponite

Laponite

0.2 2.0 0.82 ± 3.27% 6.80 ± 1.03% 3.5.2. Heterogenous Photocatalysis with TiO2, Barasym and Laponite Table 3 shows pseudo orders and half-lifetime for the photodegradation of a 21.2 μM (5 mg/L) CBZ solution, under UVC irradiation, with 0–1 mg L−1 of commercial catalytic grade TiO2, Barasym SSM-100 (synthetic montmorillonite) or Laponite® (synthetic hectorite) as heterogeneous catalysts. Synthetic clay minerals were chosen to avoid impurities present in natural minerals. High quality catalytic grade TiO2 was chosen as a "gold standard" since it is widely used for the photodegradation of organic refractory pollutants [89]. Table 3 shows pseudo orders and half-lifetime for the photodegradation of a 21.2 µM (5 mg/L) CBZ solution, under UVC irradiation, with 0–1 mg L−<sup>1</sup> of commercial catalytic grade TiO2, Barasym SSM-100 (synthetic montmorillonite) or Laponite® (synthetic hectorite) as heterogeneous catalysts. Synthetic clay minerals were chosen to avoid impurities present in natural minerals. High quality catalytic grade TiO<sup>2</sup> was chosen as a "gold standard" since it is widely used for the photodegradation of organic refractory pollutants [89]. In most heterogeneous photocatalysis studies, the catalyst concentration is from tens to thousands mg L−<sup>1</sup> [90]. We chose to test relatively low concentrations of 0.2, 0.4 and 1 mg L−<sup>1</sup> , based on our previous studies with BPS and ofloxacin [51,68]. It can be seen that the clay minerals when added alone have almost no effect (Table 3). TiO<sup>2</sup> indeed has some photocatalytic

effect, but even at a 1 mg L−<sup>1</sup> is not very effective (*t*1/2 = 121 min, n = 0). Previous studies dealing with CBZ photodegradation used three orders of magnitude higher concentrations of TiO<sup>2</sup> as a heterogeneous catalyst and obtained half-life times of 10–20 min — considerably shorter than the present study [91,92]. L−1) and (b) low (0.5 mg L−1). At 2 mg L−1 H2O2 concentration (results summarized in Table 3) the homogeneous catalyst already yields low *t*1/2 values of less than 8 min. The addition of clays as heterogeneous catalysts makes almost no difference. As for TiO2, it should be emphasized that when added alone at a low concentration (0.2 mg L−1) almost no degradation is observed

In previous studies [51,68], we have shown that a combination of low concentrations of both heterogeneous and homogeneous catalysts may yield synergistic effects and speed up the photodegradation of priority pollutants such as BPS or ofloxacin. To test this effect on CBZ, we performed photodegradation experiments of a 21.2 μM (5 mg L−1) CBZ solution (Figure 6), under UVC irradiation, with a low concentration (0.2 mg L−1) of the heterogeneous catalysts used in Section 3.5.2, at two hydrogen peroxide levels: (a) high (2.0 mg

In most heterogeneous photocatalysis studies, the catalyst concentration is from tens to thousands mg L−1 [90]. We chose to test relatively low concentrations of 0.2, 0.4 and 1 mg L−1, based on our previous studies with BPS and ofloxacin [51,68]. It can be seen that the clay minerals when added alone have almost no effect (Table 3). TiO2 indeed has some photocatalytic effect, but even at a 1 mg L−1 is not very effective (*t*1/2 = 121 min, n = 0). Previous studies dealing with CBZ photodegradation used three orders of magnitude higher concentrations of TiO2 as a heterogeneous catalyst and obtained half-life times of

#### 3.5.3. Hetero-Homogeneous Photocatalysis (Table 3), but its addition to 2.0 mg L−1 H2O2 slightly speeds up the process (*t*1/2 changes

3.5.3. Hetero-Homogeneous Photocatalysis

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10–20 min — considerably shorter than the present study [91,92].

In previous studies [51,68], we have shown that a combination of low concentrations of both heterogeneous and homogeneous catalysts may yield synergistic effects and speed up the photodegradation of priority pollutants such as BPS or ofloxacin. To test this effect on CBZ, we performed photodegradation experiments of a 21.2 µM (5 mg L−<sup>1</sup> ) CBZ solution (Figure 6), under UVC irradiation, with a low concentration (0.2 mg L−<sup>1</sup> ) of the heterogeneous catalysts used in Section 3.5.2, at two hydrogen peroxide levels: (a) high (2.0 mg L−<sup>1</sup> ) and (b) low (0.5 mg L−<sup>1</sup> ). from 6.4 to 5.9), and a small change in the pseudo-order is observed. At a low homogeneous catalyst concentration of 0.5 mg L−1, the influence of all heterogeneous catalysts is significant (Table 3): While with no heterogeneous catalyst *t*1/2 at that hydrogen peroxide amount is about 219 min, the addition of 0.2 mg L−1 of TiO2, barasym or laponite lowers *t*1/2 to 68.0, 33.2 and 37.0 min, respectively. The pseudo-order also changes, especially for the clay minerals.

**Figure 6.** Photodegradation of a 21.2 μM (5 mg L<sup>−</sup>1) carbamazepine under UVC irradiation, with 0.5 mg L<sup>−</sup>1 H2O2 alone or combined with 0.2 mg L−1 TiO2, barasym or laponite. **Figure 6.** Photodegradation of a 21.2 µM (5 mg L−<sup>1</sup> ) carbamazepine under UVC irradiation, with 0.5 mg L−<sup>1</sup> H2O<sup>2</sup> alone or combined with 0.2 mg L−<sup>1</sup> TiO<sup>2</sup> , barasym or laponite.

**4. Conclusions**  CBZ removal at relatively large (>1 mM) concentrations in industrial effluents and nanofiltration brines, or at very low (<10 μM) concentrations, should be removed in order to enable water reuse. Montmorillonite clays and organoclays may provide an efficient solution by adsorption for industrial effluents containing relatively high carbamazepine concentrations. Batch experiments have shown different adsorption capabilities of car-At 2 mg L−<sup>1</sup> H2O<sup>2</sup> concentration (results summarized in Table 3) the homogeneous catalyst already yields low *t*1/2 values of less than 8 min. The addition of clays as heterogeneous catalysts makes almost no difference. As for TiO2, it should be emphasized that when added alone at a low concentration (0.2 mg L−<sup>1</sup> ) almost no degradation is observed (Table 3), but its addition to 2.0 mg L−<sup>1</sup> H2O<sup>2</sup> slightly speeds up the process (*t*1/2 changes from 6.4 to 5.9), and a small change in the pseudo-order is observed.

bamazepine on various montmorillonite-based adsorbents, and in raw bentonite it might reach 0.5 mmole g−1. The adsorption process was confirmed by ATR-FTIR analysis on the At a low homogeneous catalyst concentration of 0.5 mg L−<sup>1</sup> , the influence of all heterogeneous catalysts is significant (Table 3): While with no heterogeneous catalyst *t*1/2 at that hydrogen peroxide amount is about 219 min, the addition of 0.2 mg L−<sup>1</sup> of TiO2, barasym or laponite lowers *t*1/2 to 68.0, 33.2 and 37.0 min, respectively. The pseudo-order also changes, especially for the clay minerals.

#### **4. Conclusions**

CBZ removal at relatively large (>1 mM) concentrations in industrial effluents and nanofiltration brines, or at very low (<10 µM) concentrations, should be removed in order to enable water reuse. Montmorillonite clays and organoclays may provide an efficient solution by adsorption for industrial effluents containing relatively high carbamazepine concentrations. Batch experiments have shown different adsorption capabilities of carbamazepine on various montmorillonite-based adsorbents, and in raw bentonite it might reach 0.5 mmole g−<sup>1</sup> . The adsorption process was confirmed by ATR-FTIR analysis on the clay particles, and CBZ desorption was not observed, validating the stability of the sorbent-CBZ complexes. While B1-bentonite organoclay exhibits high affinity at low concentrations, an S-shape isotherm was observed for the raw clays, indicating low affinity at low adsorbed CBZ. However, the maximum capacity of raw montmorillonites is higher than for organoclay. In raw montmorillonites apparently, the initial coverage of the surface

with CBZ molecules promotes enhanced adsorption of additional molecules probably by π-π interactions. According to the results, CBZ adsorption occurred only on clay surfaces, and the pollutant does not enter the internal pores of acicular clays as sepiolite. The high adsorption capability to the smectite raw clays on one hand and the low affinity at low concentrations, on the other hand, may provide a good solution for high concentrated contaminated solutions such as pharmaceutical wastewater or nano/microfiltration brine. Moreover, a combined implementation of raw montmorillonite clay and bent-B1 organoclay together can offer the advantages of raw clay for the high concentration and organoclay for the low carbamazepine concentrations.

As for AOP processes, it should be emphasized that since such processes are very specific, the description hereby focuses on the conditions of this study as for radiation intensity and CBZ concentrations. This study focuses on a relatively high initial concentration for municipal wastewater, even though there are several reports reaching such levels [7,93]. Results indicate that direct photolysis with no additional catalyst does not yield CBZ degradation. The addition of hydrogen peroxide as homogeneous photocatalysts is very effective, but only above certain levels. (H2O<sup>2</sup> > 1 mg L−<sup>1</sup> ). The heterogeneous catalysts tested (including the catalytic grade TiO2) at concentrations ranging 0–1 mg L−<sup>1</sup> do not yield almost any CBZ degradation. Combined with high concentrations of H2O<sup>2</sup> (2.0 mg L−<sup>1</sup> ) very effective degradation is observed, but this is mostly due to the homogeneous process. However, when combining a low amount of heterogeneous catalysts with very low H2O<sup>2</sup> amounts (0.5 mg L−<sup>1</sup> ) a synergistic effect is observed, and two treatments that each of them by itself are completely ineffective lead to a relatively effective process. The advantage of low amounts of catalysts is obvious, considering that the reuse of water will require the removal of the remaining catalysts—both H2O2, and clays or TiO2. It is worthwhile to emphasize that the by-products in the photodegradation process were not measured and identified at this current research stage, although previous studies have shown their presence during CBZ degradation [94,95]. Additional study is required in order to evaluate the by-products and will be conducted in the near future using LCMS-MS.

The search for more effective water techniques is driving researchers toward AOPs. However, we should recall the limitations of AOPs in general and photocatalytic water treatment devices in particular: Since processes are very specific, the challenge of dealing efficiently with multiple low-concentration priority pollutants is far from being achieved. E.L. Cates [96] strongly criticizes the pursue for "new applications, improved catalysts, and reaction mechanisms", and summarizes that "it is time to stop patting ourselves on the back for laboratory 'successes' that clearly turn a blind eye on fatal implementation hurdles". In this sense, probably a combination of processes such as adsorption on specifically tailored sorbents followed by advanced oxidation devices (or vice versa) may yield more efficient and feasible water treatment technologies for both industrial and municipal wastewater.

**Supplementary Materials:** The following supporting information can be downloaded at: https: //www.mdpi.com/article/10.3390/w14132047/s1.

**Author Contributions:** Conceptualization, G.R.; methodology, G.R., I.L. and Y.S.; software, I.L. and Y.S.; validation, G.R., I.L. and Y.S.; formal analysis, G.R., I.L. and Y.S.; investigation, G.R., I.L. and Y.S.; resources, G.R.; data curation, I.L. and Y.S.; writing—original draft preparation, G.R., I.L. and Y.S.; writing—review and editing, I.L. and G.R.; visualization, G.R., I.L. and Y.S.; supervision, G.R.; project administration, G.R.; funding acquisition, G.R. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was partially funded by CSO-MOH (Israeli Ministry of Health), in the frame of the collaborative international consortium (REWA) financed under the 2020 AquaticPollutants Joint call of the AquaticPollutants ERA-NET Cofund (GA Nº 869178).

**Institutional Review Board Statement:** Not applicable.

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Additional details on the raw data can be obtained by contacting the authors.

**Acknowledgments:** The authors would like to thank the European Commission and AKA (Finland), CSO-MOH (Israel), IFD (Denmark) and WRC (South Africa) for funding in the frame REWA international consortium (additional details in "funding" paragraph). REWA is an integral part of the activities developed by the Water, Oceans and AMR JPIs. The authors are also thankful to Chen Barak for all the technical support, Sara Azerrad (from the Shamir Research Institute) for the HPLC confirmation measurements, and the whole team of the Hydrogeology and Examination of Soil Fertility Lab at MIGAL Research Institute.

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

#### **References**

