Environmental Photocatalytic Degradation of Antidepressants with Solar Radiation: Kinetics, Mineralization, and Toxicity

This work is focused on the kinetics, mineralization, and toxicological assessments of the antidepressant drug amitriptyline hydrochloride (AMI) in UV or solar illuminated aqueous suspensions of ZnO, TiO2 Degussa P25, and TiO2 Hombikat. ZnO was proven to be the most effective photocatalyst, and it was used for all further experiments under solar irradiation. The highest reaction rate was observed at 1.0 mg/mL of catalyst loading. In the investigated initial concentration range (0.0075–0.3000 mmol/L), the degradation rate of AMI increased with the increase of initial concentration in the investigated range. The effects of H2O2, (NH4)2S2O8, and KBrO3, acting as electron acceptors, along with molecular oxygen were also studied. By studying the effects of ethanol and NaI as a hydroxyl radical and hole scavenger, respectively, it was shown that the heterogeneous catalysis takes place mainly via free hydroxyl radicals. In the mineralization study, AMI photocatalytic degradation resulted in ~30% of total organic carbon (TOC) decrease after 240 min of irradiation; acetate and formate were produced as the organic intermediates; NH4+, NO3−, NO2− ions were detected as nitrogen byproducts. Toxicity assessment using different mammalian cell lines, showed that H-4-II-E was the most sensitive one.


Introduction
Nowadays, pharmaceutically active compounds (PACs) represent a group of chemicals that are used in large quantities throughout the world. After use, PACs are excreted unchanged or as metabolites in the urine and feces, and enter the wastewater treatment plants from which some of the components are practically unmodified and discharged into the environment [1,2]. In addition, PACs can reach the environment due to improper waste disposal of unused or expired drugs [3]. Their persistence in aquatic systems can lead to a potential risk for living organisms. Among the various types of conventional methods for the purification of water, advanced oxidation processes (AOPs) have proven to be very effective for the removal of PACs from aquatic environments [4]. AOPs are defined as processes that generate highly reactive species, mostly hydroxyl radicals ( • OH radicals), which can mineralize a large number of harmful substances [5].
In recent years, a number of studies in the field of heterogeneous photocatalysis have been conducted due to the high efficiency of the degradation and mineralization of organic compounds, but also because of possible uses in the visible and UV range [6]. In most cases, heterogeneous photocatalysis is related to semiconductor photocatalysis, which can be used for the removal of organic and inorganic pollutants from the water and the gas

Photodegradation Procedure
Photodegradation experiments were performed in the photochemical cel previously by our group [19] and operational variables applied during the pho degradation experiments were shown in Table 2. Experiment of direct phot performed under the same conditions as heterogeneous photocatalytic degrad Absorption spectrum of amitriptyline hydrochloride (AMI) aqueous solution (0.0300 mmol/L). Inset represents structural formula of AMI. ZnO (99.9%, Sigma-Aldrich, crystallite size of 41.0 ± 0.9 nm, specific pore volume of 0.016 cm 3 /g, and specific surface area of 6.5 m 2 /g) [17], TiO 2 -D (75% anatase and 25% rutile, with average particle size from about 20 nm, according to the producer's specification, total pore volume 0.134 cm 3 /g, and specific surface area of 53.2 m 2 /g) [18], and TiO 2 -H (anatase, Sigma-Aldrich, specific surface area of 35-65 m 2 /g) were used as photocatalysts.

Photodegradation Procedure
Photodegradation experiments were performed in the photochemical cell described previously by our group [19] and operational variables applied during the photocatalytic degradation experiments were shown in Table 2. Experiment of direct photolysis was performed under the same conditions as heterogeneous photocatalytic degradation, but without the addition of a photocatalyst. In the investigation of the influence of electron acceptors, apart from constant streaming of O 2 , H 2 O 2 , KBrO 3 , and (NH 4 ) 2 S 2 O 8 were added to the investigated suspension. In addition, to examine the influence of active species such as • OH radicals and photogenerated holes, ethanol or NaI were added into the suspension.

Analytical Monitoring of Photodegradation
To monitor the photodegradation of AMI, high pressure liquid chromatography with a diode array detector (UFLC-DAD, Shimadzu Nexera, Tokyo, Japan) was used. Aliquots of the reaction mixture were taken before the start of irradiation and at specific time intervals during the irradiation (volume variation ca. 10%). All samples with photocatalyst were filtered through a Millipore (Millex-GV, Burlington, MA, USA, 0.22 µm) membrane filter in order to separate the catalyst particles. Prepared aliquots were analyzed on UFLC-DAD as described previously [20]. pH was measured using a combined glass electrode (pH-Electrode SenTix 20, WTW, Thermo Fisher Scientific, Waltham, MA, USA) connected to a pH meter (pH/Cond 340i, WTW). Samples for measurements of temporal changes in the total organic carbon (TOC) were irradiated at different time intervals and analyzed after filtration on an Elementar Liqui TOC II analyzer (Elementar, Langenselbold, Germany) according to Standard US EPA Method 9060A. Ion chromatography (IC) analysis was performed on a Dionex ICS 3000 Reagent-Free IC system (Thermo Scientific, Carlsbad, CA, USA) with a conductometric detector [21].

Toxicity
Assessment of the cytotoxic effect on the growth of mammalian cell lines was performed, similar to our previous investigations [22] with the difference being that in this work, the concentration of AMI was 0.0300 mmol/L and photocatalyst loading was 1.0 mg/mL. Aliquots of 2 mL suspension of AMI were taken at the beginning of the experiment, as well as at different times during the irradiation, and after that were filtered through 0.22 µm membrane filters (Sartorius, Goettingen, Germany). The cell lines were grown in DMEM medium, supplemented with 10% heat inactivated FCS, 100 IU/mL of penicillin, 100 µg/mL of streptomycin, and 0.25 µg/mL of amphotericin B. Cells were cultured in 25 mL flasks (Corning, New York, NY, USA) at 37 • C in the atmosphere of 5% CO 2 and high humidity, and sub-cultured twice a week. A single cell suspension was obtained using 0.1% trypsin with 0.04% EDTA.
Reaction mixtures of AMI and ZnO (20 µL) were added to 180 µL of the culture medium with cells. The same volume (20 µL) of ultrapure water was added to the control wells. Thus, the final concentration of all substrates was 3 µmol/L. The blank tests were performed using pure AMI solution, as well as the aqueous suspension of ZnO (without substrate), which were sonicated in the dark for 15 min and filtered through 0.22 µm membrane filters. Evaluation of cell growth was done by the colorimetric SRB assay of [23], which was modified by Cetojevic-Simin et al. [24].

Photocatalysts Screening
The efficiency of photocatalysts ZnO, TiO 2 -D, and TiO 2 -H in photocatalytic degradation of AMI under solar irradiation was investigated. As can be seen from Figure 2, photocatalytic degradation of AMI using ZnO proceeded faster than the other two photocatalysts, and after 60 min of irradiation, 94.3% of AMI was degraded. On the other hand, the degradation efficiency of AMI differed to a lesser extent in the presence of TiO 2 -D and TiO 2 -H, whereby the system with TiO 2 -H proved to be somewhat more efficient. Namely, in the presence of TiO 2 -D and TiO 2 -H, 55.3% and 72.4% of AMI was respectively degraded after 60 min of irradiation. The higher photocatalytic efficiency of AMI in the presence of ZnO could be attributed to the better mobility, generation, and separation of e − −h + pairs of ZnO in comparison with TiO 2 [25]. The blank experiments for either irradiated aqueous AMI solution (direct photolysis, Figure 2) or AMI solution in the dark showed that both the photocatalyst and irradiation are necessary for the removal of this antidepressant drug, mainly referring to the study of AMI stability in the dark, where it showed complete stability over a period longer than 850 days. In addition, Figure 2 shows the occurrence of adsorption of AMI in the dark, whereby the percentage of adsorption depended on the photocatalyst type. Higher adsorption of AMI (24.8%) was noticed with TiO 2 -H and after 30 min of sonification in the dark, as opposed to 19.1% of AMI with TiO 2 -D, while adsorption of AMI on the ZnO surface was not detected. Considering that the ZnO has been proven to be the most effective photocatalyst for the photodegradation of AMI, it was further used for all experiments. membrane filters. Evaluation of cell growth was done by the colorimetric [23], which was modified by Cetojevic-Simin et al. [24].

Photocatalysts Screening
The efficiency of photocatalysts ZnO, TiO2-D, and TiO2-H in photoca dation of AMI under solar irradiation was investigated. As can be seen f photocatalytic degradation of AMI using ZnO proceeded faster than the o tocatalysts, and after 60 min of irradiation, 94.3% of AMI was degraded. hand, the degradation efficiency of AMI differed to a lesser extent in th TiO2-D and TiO2-H, whereby the system with TiO2-H proved to be somew cient. Namely, in the presence of TiO2-D and TiO2-H, 55.3% and 72.4% of spectively degraded after 60 min of irradiation. The higher photocatalytic AMI in the presence of ZnO could be attributed to the better mobility, ge separation of e − −h + pairs of ZnO in comparison with TiO2 [25]. The blank ex either irradiated aqueous AMI solution (direct photolysis, Figure 2) or AM the dark showed that both the photocatalyst and irradiation are necessa moval of this antidepressant drug, mainly referring to the study of AMI s dark, where it showed complete stability over a period longer than 850 day Figure 2 shows the occurrence of adsorption of AMI in the dark, whereby t of adsorption depended on the photocatalyst type. Higher adsorption of was noticed with TiO2-H and after 30 min of sonification in the dark, as opp of AMI with TiO2-D, while adsorption of AMI on the ZnO surface was Considering that the ZnO has been proven to be the most effective photoc photodegradation of AMI, it was further used for all experiments.

Various Sources of Radiation
In order to investigate the influence of different types of radiation on of AMI removal, an aqueous suspension of AMI was irradiated with UV o tion. As shown in Figure 3, UV light led to the complete degradation of AM of irradiation, while after 60 min under solar radiation, 94.3% of AMI wa Table 3 are presented the kinetics parameters, and it can be seen that in th the degradation rate was 4.3 times higher than in the case of the system u diation. These results could be explained by the fact that in the reaction syst

Various Sources of Radiation
In order to investigate the influence of different types of radiation on the efficiency of AMI removal, an aqueous suspension of AMI was irradiated with UV or solar radiation. As shown in Figure 3, UV light led to the complete degradation of AMI after 30 min of irradiation, while after 60 min under solar radiation, 94.3% of AMI was removed. In Table 3 are presented the kinetics parameters, and it can be seen that in the case of UV, the degradation rate was 4.3 times higher than in the case of the system under solar radiation.
These results could be explained by the fact that in the reaction system with solar radiation, a smaller number of photons from the UV-spectrum were present, and hence the formation of highly reactive species was smaller, and accordingly, the catalytic activity of ZnO under solar radiation was also smaller. Owing to the fact that the efficiency of degradation of AMI was high enough, all further measurements were performed using solar radiation.
anomaterials 2021, 11,632 degradation of AMI was high enough, all further measurements were per solar radiation.  Table 3. Effect of radiation type on AMI degradation rate (0.0300 mmol/L) in the pr (1.0 mg/mL).

Type of Radiation
R a × 10 6 (mol/(L min)) a r b UV 11.8 0.99 Solar 2.76 0.99 R a -Degradation rate determined for the first 10 min of irradiation; r b -linear regress

Effects of ZnO Loadings
It is known that the photocatalytic degradation rate proportionally d photocatalyst loading in such a way that with increased loading, degrad creases until the optimum value. Namely, loading increase has two oppo the efficiency of the photocatalytic process. On one hand, its higher concen reaction mixture increase the number of active sites for adsorption of pol photocatalyst surface, which increases the efficiency of degradation. On th higher photocatalyst loading leads to aggregation of its particles, which dispersion of light from the photocatalyst surface. In addition, aggregat reduction in the surface area. Therefore, the photocatalyst surface is not av generation of e − −h + pairs, which can reduce the efficiency of photocatalyti [26]. A series of experiments were carried out to investigate the effect of Zn the efficiency of AMI photodegradation. ZnO loading was varied from 0.1 (Figure 4, inset). Figure 4 represents the dependence of AMI photodegrad the function of the ZnO loadings. Based on the obtained results, it can be the degradation rate increases with increased photocatalyst loadings up un and above this value the degradation rate slightly drops. Based on the ab concluded that the optimal ZnO loading for the degradation of AMI w  Table 3. Effect of radiation type on AMI degradation rate (0.0300 mmol/L) in the presence of ZnO (1.0 mg/mL).

Type of Radiation
R a × 10 6 (mol/(L min)) a r b UV 11.8 0.998 Solar 2.76 0.999 R a -Degradation rate determined for the first 10 min of irradiation; r b -linear regression coefficient.

Effects of ZnO Loadings
It is known that the photocatalytic degradation rate proportionally depends of the photocatalyst loading in such a way that with increased loading, degradation rate increases until the optimum value. Namely, loading increase has two opposite effects on the efficiency of the photocatalytic process. On one hand, its higher concentrations in the reaction mixture increase the number of active sites for adsorption of pollutants on the photocatalyst surface, which increases the efficiency of degradation. On the other hand, higher photocatalyst loading leads to aggregation of its particles, which increases the dispersion of light from the photocatalyst surface. In addition, aggregation leads to a reduction in the surface area. Therefore, the photocatalyst surface is not available for the generation of e − −h + pairs, which can reduce the efficiency of photocatalytic degradation [26]. A series of experiments were carried out to investigate the effect of ZnO loadings on the efficiency of AMI photodegradation. ZnO loading was varied from 0.1 to 2.0 mg/mL (Figure 4, inset). Figure 4 represents the dependence of AMI photodegradation rate on the function of the ZnO loadings. Based on the obtained results, it can be observed that the degradation rate increases with increased photocatalyst loadings up until 1.0 mg/mL, and above this value the degradation rate slightly drops. Based on the above, it can be concluded that the optimal ZnO loading for the degradation of AMI was 1.0 mg/mL under solar radiation. Similar results were also obtained by other researchers [27,28].

Effect of Amitriptyline Concentration
The influence of various initial concentrations of AMI on the efficie catalytic degradation was investigated. Namely, AMI concentration was range of 0.0075 to 0.3000 mmol/L ( Figure 5, inset). Figure 5 represents the degradation rate of AMI as the function of its initial concentrations, and vealed that the degradation rate increased with increases in AMI conce phenomenon is probably due to the fact that in the system with no change lyst loadings and radiation intensity, with increasing of the substrate con optimal value, the number of occupied sites on the photocatalyst is increas tion consequently increases the utilization of radiation, which in the end l crease in the photocatalytic degradation rate of AMI, which is in accordan ture data [29]. However, the increase in degradation rate is not exactly prop increase of the AMI concentration, which can be seen from the appropriat rates showed in Figure 5.

Effect of Amitriptyline Concentration
The influence of various initial concentrations of AMI on the efficiency of photocatalytic degradation was investigated. Namely, AMI concentration was varied in the range of 0.0075 to 0.3000 mmol/L ( Figure 5, inset). Figure 5 represents the photocatalytic degradation rate of AMI as the function of its initial concentrations, and the results revealed that the degradation rate increased with increases in AMI concentration. This phenomenon is probably due to the fact that in the system with no change in photocatalyst loadings and radiation intensity, with increasing of the substrate concentration to optimal value, the number of occupied sites on the photocatalyst is increased. This situation consequently increases the utilization of radiation, which in the end leads to an increase in the photocatalytic degradation rate of AMI, which is in accordance with literature data [29]. However, the increase in degradation rate is not exactly proportional to the increase of the AMI concentration, which can be seen from the appropriate degradation rates showed in Figure 5.

Effect of Amitriptyline Concentration
The influence of various initial concentrations of AMI on the efficie catalytic degradation was investigated. Namely, AMI concentration was range of 0.0075 to 0.3000 mmol/L ( Figure 5, inset). Figure 5 represents the degradation rate of AMI as the function of its initial concentrations, and vealed that the degradation rate increased with increases in AMI conce phenomenon is probably due to the fact that in the system with no change lyst loadings and radiation intensity, with increasing of the substrate co optimal value, the number of occupied sites on the photocatalyst is increas tion consequently increases the utilization of radiation, which in the end l crease in the photocatalytic degradation rate of AMI, which is in accordan ture data [29]. However, the increase in degradation rate is not exactly prop increase of the AMI concentration, which can be seen from the appropriat rates showed in Figure 5.

Effect of the Presence of Various Electron Acceptors
The recombination of e − −h + pairs represents one of the biggest problems in the application of photocatalysts. This phenomenon is very pronounced in the absence of a suitable electron acceptor, potentially causing reduction in the efficiency of the photocatalytic process. In order to increase the formation of • OH radicals and prevent recombination of e − −h + in the reaction mixture, an electron acceptor was added. O 2 is the most commonly used electron acceptor. Beside O 2 , KBrO 3 , H 2 O 2 , and (NH 4 ) 2 S 2 O 8 are usually used as electron acceptors [30]. Apart from the influence of O 2 , which was introduced in all cases, the influence of the presence of H 2 O 2 , KBrO 3 , and (NH 4 ) 2 S 2 O 8 on AMI degradation efficiency was tested ( Figure 6). anomaterials 2021, 11,632 commonly used electron acceptor. Beside O2, KBrO3, H2O2, and (NH4)2S2O used as electron acceptors [30]. Apart from the influence of O2, which was all cases, the influence of the presence of H2O2, KBrO3, and (NH4)2S2O8 on tion efficiency was tested ( Figure 6). As can be seen in Figure 6, the addition of H2O2 in the reaction system efficiency of AMI degradation as compared with the systems in which on troduced. Such results can be a consequence of formation of peroxo radical a lower activity in the process of degradation as compared to the • OH rad sides the addition of H2O2, (NH4)2S2O8 also reduced degradation efficiency pared to the systems with only O2 in the role of electron acceptor. Howev tion in degradation efficiency was only observed during the first 30 min tion, while after 60 min it was the same as with O2. On the other hand, th KBrO3 slightly elevated AMI's degradation rate as compared to the syst tained only O2, and after 60 min of irradiation, 95.8% of AMI was degraded can be explained in a way that − 3 BrO ions react with the electrons from t band of the photocatalyst, which reduces the possibility of recombining e − fore extends the lifetime of holes formed in the valence band [30]. Based on results, it can be concluded that the studied electron acceptors (H2O2, (N KBrO3) have different effects on the process of photocatalytic degradation o

Active Species Identification
The most common reaction represented in photocatalytic degradati tween the substrate and the adsorbed • OH radicals on the surface of the ph As can be seen in Figure 6, the addition of H 2 O 2 in the reaction systems reduced the efficiency of AMI degradation as compared with the systems in which only O 2 was introduced. Such results can be a consequence of formation of peroxo radicals, which have a lower activity in the process of degradation as compared to the • OH radicals [31]. Besides the addition of H 2 O 2 , (NH 4 ) 2 S 2 O 8 also reduced degradation efficiency of AMI compared to the systems with only O 2 in the role of electron acceptor. However, this reduction in degradation efficiency was only observed during the first 30 min of the irradiation, while after 60 min it was the same as with O 2 . On the other hand, the presence of KBrO 3 slightly elevated AMI's degradation rate as compared to the systems that contained only O 2 , and after 60 min of irradiation, 95.8% of AMI was degraded. These results can be explained in a way that BrO − 3 ions react with the electrons from the conduction band of the photocatalyst, which reduces the possibility of recombining e − −h + and therefore extends the lifetime of holes formed in the valence band [30]. Based on the obtained results, it can be concluded that the studied electron acceptors (H 2 O 2 , (NH 4 ) 2 S 2 O 8 , and KBrO 3 ) have different effects on the process of photocatalytic degradation of AMI.

Active Species Identification
The most common reaction represented in photocatalytic degradation is that between the substrate and the adsorbed • OH radicals on the surface of the photocatalyst or the reaction of direct charge transfer between the substrate and holes formed in the valence band of the photocatalyst [32]. To investigate the possible mechanism of photocatalytic degradation of AMI, the influence of active species such as • OH radicals and photogen-Nanomaterials 2021, 11, 632 9 of 14 erated holes in the process of photocatalytic degradation was studied by addition of the specific scavengers.
The experiments were performed by the addition of ethanol and NaI in the reaction mixture. In fact, ethanol is mainly used as a scavenger of • OH radicals [27], whereas the I − ion is used as a scavenger of adsorbed • OH radicals and photogenerated holes [32]. The obtained results (Figure 7) showed that the efficiency of AMI degradation was significantly reduced with the addition of ethanol into the reaction mixture in comparison with the results in its absence. After 60 min of irradiation in the presence of ethanol, only 41.1% of AMI was removed, whereas in the absence of ethanol, 94.3% of AMI was degraded. In addition, with the addition of NaI, the efficiency of photocatalytic degradation was reduced, whereas after 60 min of irradiation, 81.6% of AMI was removed. Based on these experimental results it can be concluded that the main photocatalytic degradation of AMI takes place via • OH radicals free in solution, which is in agreement with other investigations [33][34][35], while the valence band holes have a secondary roles, as this was also expected from other research data [27,28,36].
anomaterials 2021, 11,632 was reduced, whereas after 60 min of irradiation, 81.6% of AMI was remo these experimental results it can be concluded that the main photocatalyt of AMI takes place via • OH radicals free in solution, which is in agreem investigations [33][34][35], while the valence band holes have a secondary rol also expected from other research data [27,28,36].

Mineralization Studies of AMI
For successful application of photocatalysis, the most important info lated to the degree of mineralization achieved during the process and t disappearance of both the parent compound and by-products [5,8]. In orde degree of mineralization during photocatalytic degradation of AMI, a d total organic carbon was estimated. In general, at low pollutant concen compounds which do not form important intermediates, complete mine reactant disappearance proceed with similar half-lives. However, at hig concentration where important intermediates occur, mineralization is slo degradation of the parent compound [37]. In addition, monitoring of ammo nitrite, acetate, and formate ions by ion chromatography provides useful d ating AMI degradation. The degree of AMI mineralization, as well as th different ions formed during photocatalytic degradation is shown in Figu the obtained results it can be seen that after 240 min of irradiation only 33.2 mineralized. In addition, unsurprisingly it can be concluded that the mine the evolution of ions were slower than the kinetics of AMI removal with Z

Mineralization Studies of AMI
For successful application of photocatalysis, the most important information is related to the degree of mineralization achieved during the process and the kinetics of disappearance of both the parent compound and by-products [5,8]. In order to assess the degree of mineralization during photocatalytic degradation of AMI, a decrease of the total organic carbon was estimated. In general, at low pollutant concentration or for compounds which do not form important intermediates, complete mineralization and reactant disappearance proceed with similar half-lives. However, at higher pollutant concentration where important intermediates occur, mineralization is slower than the degradation of the parent compound [37]. In addition, monitoring of ammonium, nitrate, nitrite, acetate, and formate ions by ion chromatography provides useful data for evaluating AMI degradation. The degree of AMI mineralization, as well as the evolution of different ions formed during photocatalytic degradation is shown in Figure 8. Based on the obtained results it can be seen that after 240 min of irradiation only 33.2% of AMI was mineralized. In addition, unsurprisingly it can be concluded that the mineralization and the evolution of ions were slower than the kinetics of AMI removal with ZnO ( Figure 2). Beside CO 2 and H 2 O, looking at the formula of AMI (Figure 1, inset), it can be expected that NH 4 + and/or NO 3 − /NO 2 − as well as acetate and formate can be formed during the photocatalytic degradation of AMI.
Nitrogen-containing molecules are mineralized into NH 4 + and mostly NO 3 − , and formed ammonium ions are relatively stable, and their proportions depend mainly on the initial oxidation of nitrogen and on the irradiation time. In addition, the concentrations of NH 4 + and NO 3 − increased with increasing irradiation time, while formation of NO 2 − ions was negligible. Further, the concentration of acetate increased up until 180 min of irradiation and then decreased, while in the case of formate, a surge was observed up until 120 min, after which it decreased.

Toxicological Assessments of AMI and Formed Intermediates
Toxicity investigations are very important for environmental ecology the case of implementation of the catalyzed irradiation method. In order t cytotoxicity of AMI, as well as the intermediates of the photocatalytic deg ZnO, in vitro growth of the four cell lines were investigated, namely H-4-I HT-29, and MRC-5.
Based on the obtained results (Figure 9), it can be seen that the grow cell lines depended on irradiation time (and intermediates formed during degradation) and type of cell lines. H-4-II-E cell lines were shown to be t tive, with growth inhibition in the range of 1.34-30.6% inhibition, and were found for the longest degradation time (240 min). Significant growth observed for MRC-5 cell lines (from 9.0 to 12.9%), whereby mild stim growth was obtained for samples at 0 min and after 30 min of irradiation (3 respectively). This was probably due to the formation of cytotoxic interme min of irradiation. In Neuro-2a cells, growth was inhibited in the range 9.84%, while only mild stimulation of cell growth (1.72%) was recorded a irradiation. In the HT-29 cells, only a mild inhibition of growth was obse from 1.34 to 5.29%. Cell growth stimulation of HT-29 cells was from 2.32 to was observed after 240 and 30 min of irradiation, respectively.

Toxicological Assessments of AMI and Formed Intermediates
Toxicity investigations are very important for environmental ecology, especially for the case of implementation of the catalyzed irradiation method. In order to evaluate the cytotoxicity of AMI, as well as the intermediates of the photocatalytic degradation with ZnO, in vitro growth of the four cell lines were investigated, namely H-4-II-E, Neuro-2a, HT-29, and MRC-5.
Based on the obtained results (Figure 9), it can be seen that the growth of selected cell lines depended on irradiation time (and intermediates formed during photocatalytic degradation) and type of cell lines. H-4-II-E cell lines were shown to be the most sensitive, with growth inhibition in the range of 1.34-30.6% inhibition, and highest values were found for the longest degradation time (240 min). Significant growth inhibition was observed for MRC-5 cell lines (from 9.0 to 12.9%), whereby mild stimulation of cell growth was obtained for samples at 0 min and after 30 min of irradiation (3.41 and 0.92%, respectively). This was probably due to the formation of cytotoxic intermediates after 60 min of irradiation. In Neuro-2a cells, growth was inhibited in the range from 1.31 to 9.84%, while only mild stimulation of cell growth (1.72%) was recorded after 30 min of irradiation. In the HT-29 cells, only a mild inhibition of growth was observed, ranging from 1.34 to 5.29%. Cell growth stimulation of HT-29 cells was from 2.32 to 9.83%, which was observed after 240 and 30 min of irradiation, respectively. respectively). This was probably due to the formation of cytotoxic intermediates after 60 min of irradiation. In Neuro-2a cells, growth was inhibited in the range from 1.31 to 9.84%, while only mild stimulation of cell growth (1.72%) was recorded after 30 min o irradiation. In the HT-29 cells, only a mild inhibition of growth was observed, ranging from 1.34 to 5.29%. Cell growth stimulation of HT-29 cells was from 2.32 to 9.83%, which was observed after 240 and 30 min of irradiation, respectively.  In addition, the influence of AMI and the blank probe on cell growth was examined ( Figure 10). It was found that AMI inhibited the growth of H-4-II-E, HT-29, and Neuro-2a cells (4.66, 7.60, and 8.91%, respectively). In MRC-5 cells, stimulation of growth was observed (0.53%). Results of a blank test (in the presence of ZnO, and absence of AMI) showed cell growth inhibition in Neuro-2a and MRC-5 cells (5.84 and 7.20%, respectively) and cell growth stimulation in H-4-II-E and HT-29 cells (1.14 and 11.8%, respectively). Although ZnO nanoparticles (NPs) have wide application, i.e., they are used in food products as additives and supplements, and in containers and packaging; in the energy sector as fuels and catalysts; in consumer electronics in semiconductors and air filtration; in pharmaceuticals; in biomedical engineering; and in drinking water, they may be toxic due to their partial dissolution. Thus, the concentrations of free Zn 2+ in ZnO NP solutions needs to be measured to understand the effects of ion dissolution at different concentrations of the particles, which is important for defining toxicity. Hence, more investigation, with standard experimental conditions, is needed to better understand ZnO NP toxicity at the cellular and physiological levels, as these NPs may enter the food chains. Environmental and human exposure due to nanomaterial residues in water, air, soil, and crops is expected to increase with exposure routes, including possible bioaccumulation in the environment and food chain [38][39][40][41].
nomaterials 2021, 11, 632 11 In addition, the influence of AMI and the blank probe on cell growth was exami ( Figure 10). It was found that AMI inhibited the growth of H-4-II-E, HT-29, and Neur cells (4.66, 7.60, and 8.91%, respectively). In MRC-5 cells, stimulation of growth was served (0.53%). Results of a blank test (in the presence of ZnO, and absence of A showed cell growth inhibition in Neuro-2a and MRC-5 cells (5.84 and 7.20%, respect ly) and cell growth stimulation in H-4-II-E and HT-29 cells (1.14 and 11.8%, respectiv Although ZnO nanoparticles (NPs) have wide application, i.e., they are used in f products as additives and supplements, and in containers and packaging; in the ene sector as fuels and catalysts; in consumer electronics in semiconductors and air filtrat in pharmaceuticals; in biomedical engineering; and in drinking water, they may be t due to their partial dissolution. Thus, the concentrations of free Zn 2+ in ZnO NP solut needs to be measured to understand the effects of ion dissolution at different concen tions of the particles, which is important for defining toxicity. Hence, more investigat with standard experimental conditions, is needed to better understand ZnO NP tox at the cellular and physiological levels, as these NPs may enter the food chains. E ronmental and human exposure due to nanomaterial residues in water, air, soil, crops is expected to increase with exposure routes, including possible bioaccumulatio the environment and food chain [38][39][40][41].
Finally, based on the obtained results, none of the samples produced cell gro inhibition higher than 50% (Figures 9 and 10), and their toxicity was substantially lo compared to the cytotoxic drugs and the known toxicant, HgCl2 [42].

Conclusions
The results of this study indicate that ZnO can efficiently catalyze the ph