**1. Introduction**

Aquatic pollution derived from chemicals that are generated during various anthropogenic activities is an environmental issue of considerable importance. The increasing use of chemicals due to the current model of life and the continued growth of the population is expected to cause greater pressure on natural ecosystems and humans in the near future. Pollution caused by a multitude of pharmaceutical compounds used by humans is an environmental problem that has attracted scientific interest. Pharmaceutical substances are a large and diverse group, usually of organic compounds, which are used in high quantities throughout the world. They are also considered to be a unique class of pollutants because of their characteristics. In many cases, their behavior and fate cannot be simulated with other categories of organic pollutants [1,2].

The main source of pharmaceutical residues in the aquatic environment is wastewater treatment plants (WWTPs), which exhibit frequently limited capacity to remove these pollutants, since most of them cannot be metabolized by microorganisms [3,4]. As a result, pharmaceutical compounds reach aquatic systems at concentration levels from ng L−<sup>1</sup> to

**Citation:** Antonopoulou, M.; Papadaki, M.; Rapti, I.; Konstantinou, I. Photocatalytic Degradation of Pharmaceutical Amisulpride Using g-C3N<sup>4</sup> Catalyst and UV-A Irradiation. *Catalysts* **2023**, *13*, 226. https://doi.org/10.3390/ catal13020226

Academic Editors: Kangle Lv and Vasile I. Parvulescu

Received: 14 November 2022 Revised: 18 December 2022 Accepted: 10 January 2023 Published: 18 January 2023

**Copyright:** © 2023 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/).

µg L−<sup>1</sup> , but even at these very low concentrations, they can cause toxicological risks on living organisms [5,6].

Psychiatric drugs are a class of pharmaceuticals often prescribed and used in a wide range of mental health problems, and their use is increasing worldwide. Up to now, many studies have been carried out to investigate the presence of psychiatric compounds in the environment. Amisulpride is a typical antipsychotic drug [7], and its occurrence in the environment has been highlighted in recent monitoring studies. For instance, in hospital WWTP effluent of Ioannina (northwestern Greece), amisulpride has been detected at concentration ranges from 102.0 ng L−<sup>1</sup> to 929.4 ng L−<sup>1</sup> [8]. In seawater of the Eastern Mediterranean Sea, amisulpride was detected in the range of <0.2–5.5 ng L−<sup>1</sup> [9]. Moreover, Gago–Ferrero at al. reported amisulpride detection in the effluent of WWTP of Athens at a concentration of 0.07 ng L−<sup>1</sup> [10]. During the first wave of COVID-19, amisulpride was quantified at 16.8 ng L−<sup>1</sup> in wastewater samples of Milano and Monza (Italy) WWTPs [11]. Amisulpride has also been identified as a persistent contaminant that can be introduced into the aquatic environment and able to reach groundwaters. Moreover, the formation of its characteristic non-biodegradable N-oxide product has been verified during its treatment by conventional methods, revealing the need for advanced treatment [12,13]. Incomplete removal of various pollutants including pharmaceuticals from conventional WWTPs clearly shows the need for applying innovative technologies [3,4]. Advanced oxidation processes (AOPs) have been proposed as a tertiary treatment for wastewater [14,15]. Among different AOPs, heterogeneous photocatalysis is a promising method for the removal of organic compounds [13]. Graphitic carbon nitride (g-C3N4), an organic semiconductor, has drawn widespread attention and become a hotspot for a wide variety of photocatalytic applications (e.g., degradation of organic pollutants, CO<sup>2</sup> and NOx reduction, hydrogen production, organic selective synthesis, and water splitting). This visible light photocatalyst is an outstanding option due to its properties such as graphene-like structure, chemical and thermal stability, good visible light absorption, and low-cost [16–19]. Previously, g-C3N<sup>4</sup> has been used in photocatalytic experiments in various aqueous matrices including real hospital wastewaters [8]. However, the application of g-C3N<sup>4</sup> catalysts for the degradation of amisulpride under various matrix parameters, the identification of transformation products (TPs), and the ecotoxicological assessment of the photocatalytic treatment have not been studied so far. For this purpose, the main goal of this study was to investigate the photocatalytic degradation of amisulpride under UV-A irradiation using g-C3N<sup>4</sup> catalyst and the identification of the TPs formed in the photocatalytic process. Furthermore, the toxicity by means of microalgae *Chlorococcum* sp. and *Dunaliella tertiolecta* was evaluated. Finally, the contribution of reactive species to the degradation mechanism was estimated.

## **2. Results and Discussion**

## *2.1. Photocatalytic Degradation Kinetics*

Before the study of amisulpride photocatalytic degradation, preliminary control experiments (adsorption and UV-A photolysis) were conducted in ultrapure water (UW), and the results are depicted in Figure 1. Adsorption experiments showed that the equilibrium succeeded at 30 min, and about 13% removal of amisulpride was observed. Photolysis under UV-A light resulted in a negligible decrease (~6.0%) of amisulpride and can be explained by the UV-Vis spectra of the pharmaceutical (Figure S1). Direct photolysis under UV-A light and the employed conditions was not expected since amisulpride does not absorb at 365 nm (maximum emission wavelength of the employed irradiation source), justifying the observed trend. Negligible photodegradation of the studied pharmaceutical under UV-A irradiation [13] and simulated solar irradiation [8] was also reported in previous studies.

vious studies.

under UV-A irradiation [13] and simulated solar irradiation [8] was also reported in pre-

**Figure 1.** Adsorption and UV-A photolysis of amisulpride in UW ([amisulpride]0 = 1 mg L−1; [g-C3N4] = 300 mg L<sup>−</sup>1). **Figure 1.** Adsorption and UV-A photolysis of amisulpride in UW ([amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N<sup>4</sup> ] = 300 mg L−<sup>1</sup> ).

.

The photocatalytic performance of g-C3N4 was evaluated under UV-A light in both UW and treated municipal wastewater (WW), and the results are presented in Figure 2. The photocatalytic study reveals that g-C3N4 can be activated with the adsorption of photon energy larger than its bandgap (~2.82 eV) and subsequently generates a photo-induced electron–hole pair (e− - h+) as well as other reactive species responsible for amisulpride degradation. The reaction kinetics of amisulpride degradation were studied in both matrices, the experimental data were fitted in pseudo-first order reaction kinetics, and the corresponding kinetics parameters (degradation rate constants, half-lives, and correlation coefficients) are reported in Table 1. High removals were achieved in both matrices (Figure 2). The degradation percentages were about 86.0% within 60 min in UW and about 67.0% within 180 min in WW. Our results are in agreement with previous photocatalytic studies in which high removal percentages (up to 98%) of amisulpride were achieved in aqueous matrices using various photocatalysts, i.e., TiO2, C-doped TiO2, g-C3N4 , and The photocatalytic performance of g-C3N<sup>4</sup> was evaluated under UV-A light in both UW and treated municipal wastewater (WW), and the results are presented in Figure 2. The photocatalytic study reveals that g-C3N<sup>4</sup> can be activated with the adsorption of photon energy larger than its bandgap (~2.82 eV) and subsequently generates a photoinduced electron–hole pair (e<sup>−</sup> - h<sup>+</sup> ) as well as other reactive species responsible for amisulpride degradation. The reaction kinetics of amisulpride degradation were studied in both matrices, the experimental data were fitted in pseudo-first order reaction kinetics, and the corresponding kinetics parameters (degradation rate constants, half-lives, and correlation coefficients) are reported in Table 1. High removals were achieved in both matrices (Figure 2). The degradation percentages were about 86.0% within 60 min in UW and about 67.0% within 180 min in WW. Our results are in agreement with previous photocatalytic studies in which high removal percentages (up to 98%) of amisulpride were achieved in aqueous matrices using various photocatalysts, i.e., TiO2, C-doped TiO2, g-C3N4, and SrTiO3/g-C3N<sup>4</sup> [8,13].

SrTiO3/g-C3N4 [8,13]. A higher degradation rate was observed in UW than WW. In particular, a rate constant of k = 3.04 × 10−2 min−1 and a half-life of t1/2 = 22.80 min were calculated for amisulpride in UW. The amisulpride degradation was significantly lower in WW (k = 0.7 × 10−<sup>2</sup> A higher degradation rate was observed in UW than WW. In particular, a rate constant of k = 3.04 <sup>×</sup> <sup>10</sup>−<sup>2</sup> min−<sup>1</sup> and a half-life of t1/2 = 22.80 min were calculated for amisulpride in UW. The amisulpride degradation was significantly lower in WW (k = 0.7 <sup>×</sup> <sup>10</sup>−<sup>2</sup> min−<sup>1</sup> and t1/2 = 99.02 min).

min−1 and t1/2 = 99.02 min). **Table 1.** Kinetic parameters (rate constants, correlation coefficients (R2), half-lives (t1/2)) of amisulpride photocatalytic degradation in UW and WW. **Matrix k × 10−2 (min−1) t1/2 (min) R2** UW 3.04 22.80 0.9420 WW 0.7 99.02 0.9809 This trend indicates the significant effect of the constitution of the water matrices on the photocatalytic performance. It is well-documented in the literature that the organic as well as inorganic content (e.g., Cl−, HCO3- , NO3- , SO42-) that co-exist in WW can affect the photocatalytic performance, acting mainly as scavenger of the reactive species [20,21]. This trend indicates the significant effect of the constitution of the water matrices on the photocatalytic performance. It is well-documented in the literature that the organic as well as inorganic content (e.g., Cl−, HCO<sup>3</sup> −, NO<sup>3</sup> −, SO<sup>4</sup> <sup>2</sup>−) that co-exist in WW can affect the photocatalytic performance, acting mainly as scavenger of the reactive species [20,21]. Moreover, the constituents of WW can be adsorbed onto a catalyst's surface leading to the change of the surface charges as well as to the reduction of the available active sites [8]. Metal ions, also present in WW, can significantly affect the photocatalytic performance. The effects of Cu2+, Zn2+, Fe3+, and Al3+ were studied, and an inhibition influence on the photocatalytic performance was observed in all cases under the studied experimental conditions (Figure S2). The addition of metal ions can affect the photocatalytic degradation through (i) the decrease of O<sup>2</sup> reduction by photogenerated conduction electrons and subsequent suppression of reactive oxygen species formation and (ii) alteration of the pollutant's adsorption [22].


**Table 1.** Kinetic parameters (rate constants, correlation coefficients (R<sup>2</sup> ), half-lives (t1/2)) of amisulpride photocatalytic degradation in UW and WW. Metal ions, also present in WW, can significantly affect the photocatalytic performance. The effects of Cu2+, Zn2+, Fe3+, and Al3+ were studied, and an inhibition influence on the photocatalytic performance was observed in all cases under the studied experimental con-

Moreover, the constituents of WW can be adsorbed onto a catalyst's surface leading to the change of the surface charges as well as to the reduction of the available active sites [8].

*Catalysts* **2023**, *13*, x FOR PEER REVIEW 4 of 14

**Figure 2.** Photocatalytic degradation of amisulpride under UVA irradiation in UW and WW ([amisulpride]0 = 1 mg L<sup>−</sup>1; [g-C3N4] = 300 mg L−1). **Figure 2.** Photocatalytic degradation of amisulpride under UVA irradiation in UW and WW ([amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N<sup>4</sup> ] = 300 mg L−<sup>1</sup> ).

The effect of initial pH was also evaluated for the degradation of amisulpride under similar conditions in WW as matrix (Figure S3). The highest removal of amisulpride was observed at natural pH (about 7.6). The pHPZC of the used catalyst was determined to be ~4.9. Below or above this value, the charge of the catalyst surface is positive and negative, respectively. Similarly, amisulpride (pKa 9.37) molecules are mainly protonated at pH ~ 7.6. Subsequently, under the experimental conditions, electrostatic attractions occur between amisulpride and negatively charged catalyst's surface also leading to the highest removal. On the other hand, under pH 4, electrostatic repulsions of positive charged catalyst, and amisulpride molecules significantly decrease the degradation. Under pH 10, some electrostatic repulsion of negative charged catalysts and partially negatively charged amisulpride molecules could also be considered that verified the slight decrease of the observed degradation. The effect of initial pH was also evaluated for the degradation of amisulpride under similar conditions in WW as matrix (Figure S3). The highest removal of amisulpride was observed at natural pH (about 7.6). The pHPZC of the used catalyst was determined to be ~4.9. Below or above this value, the charge of the catalyst surface is positive and negative, respectively. Similarly, amisulpride (pKa 9.37) molecules are mainly protonated at pH ~ 7.6. Subsequently, under the experimental conditions, electrostatic attractions occur between amisulpride and negatively charged catalyst's surface also leading to the highest removal. On the other hand, under pH 4, electrostatic repulsions of positive charged catalyst, and amisulpride molecules significantly decrease the degradation. Under pH 10,some electrostatic repulsion of negative charged catalysts and partially negatively charged amisulpride molecules could also be considered that verified the slight decrease of the observed degradation.

#### *2.2. Role of Reactive Species to the Degradation Mechanism 2.2. Role of Reactive Species to the Degradation Mechanism*

The contribution of HO•, h+,, and O2•- was evaluated using isopropanol (i-PrOH), methanol (MeOH), and p-benzoquinone (p-BQ), respectively. The scavengers were selected due to their high-rate constant reaction with the corresponding species. More specifically, i-PrOH is a well-documented HO• scavenger presenting a high-rate constant reaction with the radical equal to 1.9 × 109 L mol−1 s−1 [23,24]. MeOH can quench both HO• and h+ and react with HO• with a rate constant of 1 × 109 L mol−1 s−1 [20,25]. Para-benzoquinone (p-BQ) can quench O2•- with a rate constant of 1.9 × 109 L mol−1 s−1 [20]. The pseudofirst order rate constants as well as the degradation profile in the presence of scavengers are presented in Table 2 and Figure 3, respectively. The contribution of HO• , h+,, and O<sup>2</sup> •− was evaluated using isopropanol (i-PrOH), methanol (MeOH), and p-benzoquinone (p-BQ), respectively. The scavengers were selected due to their high-rate constant reaction with the corresponding species. More specifically, i-PrOH is a well-documented HO• scavenger presenting a high-rate constant reaction with the radical equal to 1.9 <sup>×</sup> <sup>10</sup><sup>9</sup> L mol−<sup>1</sup> s −1 [23,24]. MeOH can quench both HO• and h<sup>+</sup> and react with HO• with a rate constant of 1 <sup>×</sup> <sup>10</sup><sup>9</sup> L mol−<sup>1</sup> s −1 [20,25]. Para-benzoquinone (p-BQ) can quench O<sup>2</sup> •− with a rate constant of 1.9 <sup>×</sup> <sup>10</sup><sup>9</sup> L mol−<sup>1</sup> s −1 [20]. The pseudo-first order rate constants as well as the degradation profile in the presence of scavengers are presented in Table 2 and Figure 3, respectively.


**Table 2.** Kinetic parameters (rate constants, correlation coefficients (R<sup>2</sup> )), and percentages of inhibition (% ∆k) of amisulpride photocatalytic degradation in the presence of scavengers. HଶOଶ Oଶ ି → HO OHି Oଶ (3)

ି Hା → HOଶ

Oଶ

2 HOଶ

**Table 2.** Kinetic parameters (rate constants, correlation coefficients (R2)), and percentages of inhibi-

**Scavenger k × 10−2 (min−1) R2 Δk (%)**

i-PrOH 1.76 0.9800 42.1 MeOH 0.21 0.9292 93.1 p-BQ 0.80 0.9720 73.7

ିcan lead to the formation of HO• through

→ Oଶ HଶOଶ (2)

(1)

Without Scavenger 3.04 0.9420 **-**

Among the employed scavengers, MeOH caused the highest inhibition (93.1%), as it can quench both HO• and h+. In contrast, the addition of i-PrOH provoked only 42.1% inhibition, highlighting the significant role of h+ and the low contribution of HO•. This is in agreement with the literature data, as during the activation of g-C3N4 photocatalysts, the formation of HO• through direct water or OH− oxidation by holes is not promoted, since VB edges of g-C3N4 catalysts are still less positive than the redox potential of OH−/HO• (1.99 eV vs. NHE) [26,27]. The significant role of h+ was further verified with experiments using acetonitrile as reaction media (data not shown). The addition of p-BQ led to an inhibition of 73.7%. Superoxide anions can directly degrade organic pollutants

tion (% Δk) of amisulpride photocatalytic degradation in the presence of scavengers.

*Catalysts* **2023**, *13*, x FOR PEER REVIEW 5 of 14

through oxidative mechanisms. In addition, Oଶ

the following reactions [27]:

**Figure 3.** Photocatalytic degradation of amisulpride in the presence of scavengers ([amisulpride]0 = 1 mg L<sup>−</sup>1; [g-C3N4] = 300 mg L−1; [i-PrOH] = 0.1 M; [MeOH]= 0.1 M; [p-BQ] = 0.2 mM, matrix: UW). **Figure 3.** Photocatalytic degradation of amisulpride in the presence of scavengers ([amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N<sup>4</sup> ] = 300 mg L−<sup>1</sup> ; [i-PrOH] = 0.1 M; [MeOH]= 0.1 M; [p-BQ] = 0.2 mM, matrix: UW).

 Among the employed scavengers, MeOH caused the highest inhibition (93.1%), as it can quench both HO• and h<sup>+</sup> . In contrast, the addition of i-PrOH provoked only 42.1% inhibition, highlighting the significant role of h<sup>+</sup> and the low contribution of HO• . This is in agreement with the literature data, as during the activation of g-C3N<sup>4</sup> photocatalysts, the formation of HO• through direct water or OH− oxidation by holes is not promoted, since VB edges of g-C3N<sup>4</sup> catalysts are still less positive than the redox potential of OH−/HO• (1.99 eV vs. NHE) [26,27]. The significant role of h<sup>+</sup> was further verified with experiments using acetonitrile as reaction media (data not shown). The addition of p-BQ led to an inhibition of 73.7%. Superoxide anions can directly degrade organic pollutants through oxidative mechanisms. In addition, O •− 2 can lead to the formation of HO• through the following reactions [27]:

$$\rm O\_2^{\bullet-} + H^+ \to HO\_2^{\bullet} \tag{1}$$

$$2\text{ H}\text{O}\_2^\bullet \rightarrow \text{O}\_2 + \text{H}\_2\text{O}\_2\tag{2}$$

$$\mathrm{H\_2O\_2 + O\_2^{\bullet -} \to HO^{\bullet} + OH^- + O\_2} \tag{3}$$

Our results are consistent with a previous study [28] that reported that h<sup>+</sup> and O<sup>2</sup> •− are the main species in the photocatalytic degradation of refractory contaminants using g-C3N4- based catalysts. According to the above analysis, the proposed photocatalytic mechanism involves the formation of electron–hole pairs in the conduction band (CB) and valence band (VB) of g-C3N4. The holes (h<sup>+</sup> ) in the VB of g-C3N<sup>4</sup> participate directly in the degradation process, whereas the e- on the CB can reduce O<sup>2</sup> to O•− 2 .

#### *2.3. Photocatalytic Degradation Mechanism 2.3. Photocatalytic Degradation Mechanism*  The TPs that are produced at the first stages of the photocatalytic processes were

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The TPs that are produced at the first stages of the photocatalytic processes were identified in both UW and WW using mass spectrometry techniques and interpretation of their mass spectra obtained in positive ionization mode. In total, three TPs were detected in both matrices, and their structural identification was based on their molecular and fragment ions (Table 3) as well as on previous literature data [13]. identified in both UW and WW using mass spectrometry techniques and interpretation of their mass spectra obtained in positive ionization mode. In total, three TPs were detected in both matrices, and their structural identification was based on their molecular and fragment ions (Table 3) as well as on previous literature data [13].

**Table 3.** Mass measurements ([M + H]<sup>+</sup> ) and molecular formula of the identified TPs. **Table 3.** Mass measurements ([M+H]+) and molecular formula of the identified TPs. **TPs [M+H]+ Molecular Formula** 


Three pathways are proposed and depicted in Figure 4. The first pathway includes the scission of the methoxy group of the amisulpride molecule and the formation of AMI1. This TP was also identified by Skibi ´nski et al., 2011 [29] during the photodegradation of the studied pharmaceutical. Similarly, it was identified during the photocatalytic treatment of pharmaceuticals mixture also containing amisulpride [13]. The other pathway proceeds through dealkylation and more specifically through the cleavage of the ethyl group linked to the N-atom (AMI 2). The last route leads to the formation of a characteristic N-oxide TP viaan oxygen transfer mechanism [24]. N-oxide amisulpride is a characteristic TP that has been identified during the treatment of amisulpride by photolysis [29] and heterogeneous photocatalysis [13]. The absence of hydroxylated TPs verifies the low contribution of HO• to the degradation mechanism, reported in the previous section. The proposed degradation routes are consistent with previous work focused on the photocatalytic degradation of pharmaceuticals mixture also containing amisulpride [13]. Three pathways are proposed and depicted in Figure 4. The first pathway includes the scission of the methoxy group of the amisulpride molecule and the formation of AMI1. This TP was also identified by Skibiński et al., 2011 [29] during the photodegradation of the studied pharmaceutical. Similarly, it was identified during the photocatalytic treatment of pharmaceuticals mixture also containing amisulpride [13]. The other pathway proceeds through dealkylation and more specifically through the cleavage of the ethyl group linked to the N-atom (AMI 2). The last route leads to the formation of a characteristic N-oxide TP viaan oxygen transfer mechanism [24]. N-oxide amisulpride is a characteristic TP that has been identified during the treatment of amisulpride by photolysis [29] and heterogeneous photocatalysis [13]. The absence of hydroxylated TPs verifies the low contribution of HO• to the degradation mechanism, reported in the previous section. The proposed degradation routes are consistent with previous work focused on the photocatalytic degradation of pharmaceuticals mixture also containing amisulpride [13].

**Figure 4.** Photocatalytic degradation pathways of amisulpride. **Figure 4.** Photocatalytic degradation pathways of amisulpride.

The formation of TPs was also followed during the photocatalytic treatment. As presented in Figure 5, the TPs in UW and WW attained their maximum area in 20 min and 60 min, respectively. Thereafter, the area of the three TPs in both matrices were readily reduced. Subsequently, the cleavage of the rings can take place, and low molecular weight TPs can be produced before complete mineralization. After prolonged application of heterogeneous photocatalysis, two low-molecular-weight carboxylic acids, i.e., formic and acetic acid, were identified. sented in Figure 5, the TPs in UW and WW attained their maximum area in 20 min and 60 min, respectively. Thereafter, the area of the three TPs in both matrices were readily reduced. Subsequently, the cleavage of the rings can take place, and low molecular weight TPs can be produced before complete mineralization. After prolonged application of heterogeneous photocatalysis, two low-molecular-weight carboxylic acids, i.e., formic and acetic acid, were identified.

The formation of TPs was also followed during the photocatalytic treatment. As pre-

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**Figure 5.** Evolution profiles of TPs identified during the photocatalytic degradation of amisulpride in: (**a**) UW; and (**b**) WW ([amisulpride]0 = 1 mg L<sup>−</sup>1; [g-C3N4] = 300 mg L−1). **Figure 5.** Evolution profiles of TPs identified during the photocatalytic degradation of amisulpride in: (**a**) UW; and (**b**) WW ([amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N<sup>4</sup> ] = 300 mg L−<sup>1</sup> ).

#### *2.4. Ecotoxicity Evolution*

The evolution of toxicity against freshwater and marine microalgae was evaluated in both matrices (Figures 6 and 7). According to the results, low toxicity was observed at 0 min of the process in UW (Figures 6a and 7a). In contrast, the untreated solution of amisulpride in WW (0 min) led to over 25% inhibition of growth rate for both microalgae at contact time of 72 h (Figures 6b and 7b). The high growth inhibition rates observed in

WW can be correlated with the composition of this matrix that can contain compounds that exert toxic effects to the exposed microalgae. This was verified by the initial toxicity of the wastewater sample which showed an inhibition of 18.1%. Similar trends were also observed by Gomes et al., 2019 [30] and Antonopoulou et al., 2016 [20]. Only a slight increase in toxicity was observed at the first stages of the photocatalytic treatment in both matrices. This increase can be correlated with the formation of the identified TPs. It is worth mentioning that the highest % growth inhibition rates were observed at 20 min and 60 min in UW and WW, respectively, where the TPs showed their maximum formation. Similar to our results, enhanced toxicity at the first stages was also noticed during the photocatalytic degradation of various contaminants in aqueous matrices by different AOPs [20,31]. However, the contribution of the TPs in the overall toxicity can be characterized as low. With the application of photocatalytic treatment, the % inhibition was significantly reduced in all cases. This clearly indicates that heterogeneous photocatalysis using g-C3N<sup>4</sup> can proceed without the formation of toxic TPs. It is also able to significantly reduce the toxicity derived from municipal wastewater. *Catalysts* **2023**, *13*, x FOR PEER REVIEW 9 of 14

**Figure 6.** Evolution of toxicity during the photocatalytic process in: (**a**) UW; and (**b**) WW using **Figure 6.** Evolution of toxicity during the photocatalytic process in: (**a**) UW; and (**b**) WW using *Chlorococcum* sp. ([amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N4] = 300 mg L−<sup>1</sup> ).

*Chlorococcum sp.* ([amisulpride]<sup>0</sup> = 1 mg L−1; [g-C3N4] = 300 mg L−1).

**Figure 7.** Evolution of toxicity during the photocatalytic process in: (**a**) UW; and (**b**) WW using *Dunaliella tertiolecta* ([amisulpride]<sup>0</sup> = 1 mg L−1; [g-C3N4] = 300 mg L−1). **Figure 7.** Evolution of toxicity during the photocatalytic process in: (**a**) UW; and (**b**) WW using *Dunaliella tertiolecta* ([amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N<sup>4</sup> ] = 300 mg L−<sup>1</sup> ).

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

#### *3.1. Chemicals and Reagents 3.1. Chemicals and Reagents*

Αmisulpride (98%) was obtained from Tokyo chemical industry CO., LTD. Graphitic carbon nitride (g-C3N4) (BET SSA 35 m2g−1, particle size of 25 nm, E<sup>g</sup> = ~2.82 eV) was used as the photocatalyst [8,32]. Acetonitrile, isopropanol, methanol, and water of HPLC-grade solvents were supplied by Fisher Chemical. *p*-benzoquinone (≥ 98%), formic acid, BG-11 medium (Cyanobacteria BG-11 Freshwater Solution), and F/2 medium (Guillard′s (F/2) Marine Water Enrichment Solution) were purchased from Sigma-Aldrich. Ultrapure water and treated wastewater (pH = 7.45 ± 0.02; conductivity = 312.81 ± 20.3 μS cm−1; total suspended solids = 1.97 ± 0.022 mg L−1; chemical oxygen demand = 17.6 ± 1.92 mg L−1; PO<sup>4</sup> = 3.85 ± 0.024 mg L−1; SO<sup>4</sup> 2− = 32.1 ± 1.56 mg L−1; NO<sup>3</sup> <sup>−</sup> = 24.2 ± 0.78 mg L−1) were used as matrices. The microalgae species *Chlorococcum sp.* (strain SAG 22.83) and *Dunaliella tertiolecta* (strain CCAP 19/6B) were purchased from the bank SAG collection of the Göttingen University (Germany) and Scottish Marine Institute, Oban, Argyll, Scotland, respectively. Amisulpride (98%) was obtained from Tokyo chemical industry Co., Ltd. (Tokyo, Japan). Graphitic carbon nitride (g-C3N4) (BET SSA 35 m2g −1 , particle size of 25 nm, E<sup>g</sup> = ~2.82 eV) was used as the photocatalyst [8,32]. Acetonitrile, isopropanol, methanol, and water of HPLC-grade solvents were supplied by Fisher Chemical (Waltham, MA, USA). *p*-benzoquinone (≥98%), formic acid, BG-11 medium (Cyanobacteria BG-11 Freshwater Solution), and F/2 medium (Guillard0 s (F/2) Marine Water Enrichment Solution) were purchased from Sigma-Aldrich (St. Louis, MO, USA). Ultrapure water and treated wastewater (pH = 7.45 <sup>±</sup> 0.02; conductivity = 312.81 <sup>±</sup> 20.3 <sup>µ</sup>S cm−<sup>1</sup> ; total suspended solids = 1.97 ± 0.022 mg L−<sup>1</sup> ; chemical oxygen demand = 17.6 <sup>±</sup> 1.92 mg L−<sup>1</sup> ; PO<sup>4</sup> <sup>3</sup><sup>−</sup> = 3.85 <sup>±</sup> 0.024 mg L−<sup>1</sup> ; SO<sup>4</sup> <sup>2</sup><sup>−</sup> = 32.1 <sup>±</sup> 1.56 mg L−<sup>1</sup> ; NO<sup>3</sup> <sup>−</sup> = 24.2 <sup>±</sup> 0.78 mg L−<sup>1</sup> ) were used as matrices. The microalgae species *Chlorococcum* sp. (strain SAG 22.83) and *Dunaliella tertiolecta* (strain CCAP 19/6B) were purchased from the bank SAG collection of the Göttingen University (Germany) and Scottish Marine Institute, Oban, Argyll, Scotland, respectively.

3−

#### *3.2. Photocatalytic Experiments*

Photocatalytic experiments were carried out on a system equipped with 4 black light tubes from Sylvania (maximum emission at 365 nm) and electric fans to avoid overheating. A cylindrical pyrex glass cell with a maximum capacity of 200 mL was used as a photoreactor. In a typical experiment, the following experimental conditions [amisulpride]<sup>0</sup> = 1 mg L−<sup>1</sup> ; [g-C3N4] = 300 mg L−<sup>1</sup> were adopted. The suspensions were kept in the dark for 30 min under continuous stirring to achieve adsorption equilibrium before starting the radiation. The photon flux (I0) entering the solution was calculated to be 1.1 <sup>±</sup> 0.02 <sup>µ</sup>Einstein s−<sup>1</sup> under the applied conditions, periodically using potassium ferrioxalate as actinometer [33,34]. PVDF 0.22 µm by Millex-GV was used to remove the catalyst particles.
