Antiepileptic drugs in aquatic environments: Occurrence, toxicity, transformation mechanisms and fate

Abstract The increasing consumption of antiepileptic drugs (AEDs) has led to their widespread presence in aquatic environments, which poses a serious threat to human health and aquatic organisms. However, there is currently lack of a review to provide basic data and research directions for future scholars. This article was carried out to summarize toxicity, analytical methods, concentrations, transformation mechanisms and fate of AEDs and their metabolites/transformation products in aquatic environments according to existing literature. AEDs and their metabolites/transformation products were widely found in surface water, groundwater and drinking water with concentrations from ng L−1-μg L−1. AEDs at environmentally relevant concentrations have affected aquatic ecosystem, and partial AEDs could commonly cause the damage of antioxidant capacity for some aquatic organisms. The transformation of AEDs was always in the direction of oxidation, and hydroxylated and ketonizated products were conductive objectives in three pathways including human body, microorganisms and water treatment processes as the difference of transformation pathways was also observed. AEDs were supposed to dominate the concentration distribution in water phase than in sediment phase due to their chemical properties. Furthermore, photolysis was a main process for AEDs in the photic zone of receiving waters. Meanwhile, the current deficiencies of AEDs research are also pointed out. It is mainly reflected in the insufficient toxicity data and incomplete transformation pathways of AEDs and their metabolites/transformation products, which will underestimate their environmental hazards. Finally, the article also pointed out that more attention should be paid to identifying transformation products combining mechanisms analysis with nontarget analysis. Graphical Abstract


Introduction
The antiepileptic drugs (AEDs), as a class of contaminants of pharmaceuticals and personal care products (PPCPs), have been frequently detected in aquatic environments (Huerta-Fontela et al., 2011;Tran et al., 2018). AEDs at environmentally relevant concentrations have attracted increasing concern because they can cause serious health effects such as teratogenic, DNA damage, and so on (Herrmann et al., 2015;Yan et al., 2021). Recently, some studies found that metabolites/ transformation products of AEDs could also form in the processes of metabolism and transformation via human body, microorganisms and water treatment (Bahlmann et al., 2014;Han et al., 2018;Kaiser et al., 2014). And partial metabolites/transformation products were pharmacologically active or genotoxic compounds, which were more toxic than their parents (Han et al., 2018;Heye et al., 2016). Thus, the coexistence of AEDs and their metabolites/transformation products is a threat to aquatic environments.
The large consumption of AEDs is the main cause of their high concentrations occurring in the environment. In clinic, AEDs are used to control the seizure of epileptic symptoms. The statistical data according to the World Health Organization (WHO) showed that there were approximately 50 million patients with epilepsy around the world, making it one of the most common neurological diseases globally (WHO, 2019). Long-term or even life-long drug usage is still the first choice for epilepsy patients (Wang and Chen, 2019). So, the annual consumption of AEDs is extremely large. For example, carbamazepine (CBZ) was consumed at 40 tons in England (Jones et al., 2002), 35 tons in USA (Zhang et al., 2008), and more than 1000 tons around the world (Naghdi et al., 2019). Furthermore, demand of AEDs in the clinical treatment of epileptic seizures increased every year. The recent research showed a 15.6% compound annual growth rate in AEDs sales along with a significant growth in manufacturing of various AEDs in Pakistan (Mogal and Aziz, 2020). In Germany, consumption of gabapentin (GAB) increased to 73.3 tons in 2012 from 58.9 tons in 2009 (Herrmann et al., 2015). Nevertheless, the ultimate destination of these AEDs after use is municipal wastewater treatment plants (WWTPs).
AEDs are excreted from the body in the form of metabolites or the original medicine. These pharmaceutical pollutants are generally metabolically stable; therefore, traditional wastewater treatment processes cannot effectively remove AEDs or their metabolites/transformation products, leading to high concentrations of these pollutants in the WWTP effluents, and sometimes even higher than concentrations in the influents (Gobel et al., 2007;Gros et al., 2010;Vieno et al., 2007). As a result, AEDs or their metabolites/transformation products are frequently discharged into aquatic environments with concentrations ranging from ng L −1 to μg L −1 (Gurke et al., 2015;Petrie et al., 2015;Verlicchi et al., 2012). For example, CBZ has the highest occurrence (59%) among AEDs with concentrations ranging from 0.1 to 2750 ng L −1 in wastewater and aquatic environments worldwide (Hernández-Tenorio, González-Juárez, et al., 2022). In addition, partial AEDs like CBZ and GAB have a long half-life and do not easily biodegrade in aquatic environments (Ben et al., 2018;Kasprzyk-Hordern et al., 2009). Therefore, the concomitant aquatic environment problems gradually have aroused much concern. However, the current knowledge about AEDs in aquatic environments mainly focuses on the parent compounds, and there is a lack of overall research on the metabolites/transformation products. Specific problems include: (i) How the metabolites/transformation products of AEDs are transformed in different systems such as human body, microorganisms and water treatment processes; (ii) Whether there are common intermediate products and final products, which determined the final forms in aquatic environment of AEDs; (iii) How toxic these products are to aquatic organisms and human body, which decides the environmental hazards of these AEDs. More attention should be paid to the research of AEDs and their metabolites/transformation products in aquatic environments.
Thus, this paper aims to: (i) Summarize toxicity, analytical methods and concentration levels of AEDs and their metabolites/transformation products; (ii) Outline the transformation pathways of AEDs in the processes of metabolizing in the human body, microorganisms and water treatment; (iii) Figure out the fate of AEDs in aquatic environments and research directions in the future.
Long-term exposure to AEDs in aquatic environments may pose a potential threat to aquatic ecosystems and human health. CBZ is ubiquitous in the aquatic environment and difficult to degrade. The current evidence implied that CBZ has a certain toxicity to aquatic organisms. Specifically, it was found that CBZ can produce an oxidative effect on Asian clams Corbicula fluminea, and CBZ at environmentally relevant concentrations exerts a negative effect on the clams, has homologous estrogenic activity and induces reproductive toxicity on Chinese rare minnows (Gobiocypris rarus) (Chen et al., 2014;. In addition, CBZ can also induce liver histopathological changes, a gender-specific response in hepatic proteome (Yan, Wang, Liang, et al., 2018) and lipid metabolism disorder (Xin et al., 2021), and cause DNA damage  of Chinese rare minnows (G. rarus). Almeida et al. also found CBZ at μg L −1 concentration levels can lead to the impairment of antioxidant enzymes compromising the neutralization of reactive oxygen species, and thus the ability to cope with oxidative stress (for clam Ruditapes philippinarum) (Almeida et al., 2015). Additionally, CBZ can also induce effects on morphology, reproduction and lipid peroxidation at the environmental concentrations in H. circumcincta (Desbiolles et al., 2020). Similarly, GAB cannot be metabolized in the human body and is excreted as the parent compound. Furthermore, GAB has teratogenic effects, is difficult to biodegrade and has a high concentration in urban rivers (Herrmann et al., 2015;Kasprzyk-Hordern et al., 2009;Prabhu et al., 2008). GAB at environmentally relevant concentrations simultaneously affects various vital functionalities including antioxidant, immune and nervous systems of zebrafish (Danio rerio) at early developmental stage (He et al., 2019). In the toxicity studies of OXC and diazepam, both of them was supposed to impact the total antioxidant capacity for H. circumcincta and Lepomis gibbosus (Brandao et al., 2013;Desbiolles et al., 2020).
In addition to the parent compounds of AEDs, their metabolites/transformation products have also attracted extensive attention from researchers. There is evidence that more than 30 CBZ metabolites, including pharmacologically active or genotoxic compounds, have been identified in humans (Kaiser et al., 2014;Leclercq et al., 2009;Yin et al., 2017). And these metabolites can have more toxic than the parent CBZ. For example, Heye et al. investigated the chronic toxicity of 10,11-epoxy-10,11-dihydro-CBZ (CBZ-EP) on the nonbiting midge Chironomus riparius. The results showed that the metabolite CBZ-EP was significantly more toxic than the parent CBZ (Heye et al., 2016). Han et al. investigated the changes of the cytotoxicity and genotoxicity for CBZ treated by the processes of chlorination, chloramination and ozonation. The transformation products mixture after chlorination and chloramination processes caused DNA damage effects in SOS/umu test. Acridine, acridone, chlorinated acridone and aridine-9,10(4aH)-dicarbaldehyde were identified and predicted to be the DNA damaging agents. Furthermore, transformation products mixture after ozonation enhanced the cytotoxicity and the chromosome damage effects on CHO-K1 cells (Han et al., 2018). In the study of Li et al., thirteen transformation products of OXC under different water disinfection processes (ozonation, chlorination, and UV irradiation) were identified, including acridine with carcinogenic properties. An increasing acute toxicity was observed during UV irradiation probably because of the accumulation of acridine Yin et al., 2017). Also, gabapentin-lactam, as a transformation product during biological processes of GAB, were suspected to be more toxic than its parent, and lead to adverse impacts on the zebrafish (D. rerio) embryos (He et al., 2021).
Therefore, AEDs and their metabolites/transformation products indeed influence the aquatic environment and organisms to a certain extent. For parent compounds, partial AEDs including  c the value was the method quantification limit.
CBZ, GAB, OXC and diazepam can commonly cause the damage of antioxidant capacity, which were paid more attention. For their metabolites/transformation products, toxicity data was still extremely deficient while nearly dozens of kinds have been identified in the treatment processes of drinking water treatment plants (DWTPs) and WWTPs. The reasons mainly included two aspects: (i) The real transformation pathways were incomplete, which mean that many unknown metabolites/transformation products were not observed by present analytical methods; (ii) The standards were unavailable as partial metabolites/transformation products have been detected. Due to the occurrence of large amounts of metabolites/transformation products, their mixture toxicity was also required to be addressed. Furthermore, there are metabolites/transformation products of only three AEDs introduced in this section. Therefore, more toxicity data including AEDs and their metabolites/transformation products need to be acquired to evaluate their environmental hazards.

Analytical methods
Generally, sample pretreatment and chemical analysis decide whether the concentration measured by this analytical method can reflect the real concentration in the environmental samples. Table  2 showed analytical methods of AEDs in various environmental samples. The concentrations of AEDs in environmental samples are at trace levels, so sample pretreatment is a key step that can affect the accuracy and sensitivity of the whole analytical method. Selections of enrichment methods and elution solvent are necessary in the processes of pretreatment to achieve the requirements of instrument detection for target analytes. The analytical methods of AEDs in different environmental samples were reported in recent years, which include urban sewage, drinking water and sludge, are shown in Table 2. The results showed that solid phase extraction (SPE) remains a preferred pretreatment method for analyzing AEDs in various environmental samples. As a robust separation and enrichment technology, SPE has the advantages of low consumption of organic solvents, short extraction time and high recovery (Dimpe and Nomngongo, 2016). In addition, SPE has high selectivity and enables simultaneous extraction and enrichment of multiple categories of organic pollutants from environmental samples by selecting and combining different extraction cartridges filling with different stationary phase adsorbents. Specifically, hydrophilic-lipophilic balanced (HLB, 6 mL, 200 mg) cartridge was not susceptible to be most commonly used for enriching AEDs in aquatic environments because of its retention for substances with a wide polarity range (Gurke et al., 2015;Huerta-Fontela et al., 2011;Kleywegt et al., 2011;Sadutto et al., 2020;Wang et al., 2011). As for elution solvent, methanol was preferred with the robust recovery (96-122%) as other combinations like methanol/water/formic acid (v/v/v = 90:10:1) and methanol/water (v/v = 1:1) were similarly used in some studies (Gurke et al., 2015;Kleywegt et al., 2011;. Entirely, existing sample pretreatment method was efficient to achieve the recovery (75-122%) of AEDs in water or sewage sludge samples (Huerta-Fontela et al., 2011;. In addition, it is often necessary to extract AEDs from the sludge prior to SPE in the detection of AEDs in sludge. The common extraction method is accelerated solvent extraction (ASE). For example, Ding et al. employed ASE to extract AEDs in sludge from wastewater treatment plants and they optimized the extract parameters, including extraction solvent, extraction temperature, extraction pressure and extraction time. Then, the extracts were cleaned up using SPE followed by determination using liquid chromatography coupled with tandem mass spectrometry (LC-MS/MS). The extraction method was successfully applied in an actual analysis of CBZ in sludge of WWTPs and the recovery rate reached 88.1% (Ding et al., 2011). Radjenović et al. also used a similar pretreatment method to detect CBZ in sludge of WWTPs in Spain and obtained the recovery rate of 85.2% (Radjenovi et al., 2009).
It can be seen from Table 2 that detection methods of AEDs in environmental samples are mostly liquid chromatography-mass spectrometry (LC-MS). To improve the detection sensitivity of AEDs, high-performance liquid chromatography-tandem mass spectrometry (HPLC-MS/MS) and ultra-high-performance liquid chromatography-tandem mass spectrometry (UPLC-MS/MS) with higher selectivity and sensitivity were developed in some reports. Gurke et al. developed an HPLC-MS/MS method for detection of AEDs including CBZ, GAB, lamotrigine, pregabalin, levetiracetam, topiramate, OXC and primidone in WWTP influents and effluents. The limit of quantification (LOQ) values ranged from 50 to 200 ng L −1 (Gurke et al., 2015). Huerta-Fontela et al. used the method of UPLC-MS/MS to detect CBZ, diazepam, lorazepam, phenytoin and primidone in WWTP influents and effluents. The detection parameters of UPLC-MS/MS were optimized. The LOQ values of the method were from 0.2 to 1.5 ng L −1 , respectively. The matrix standard recovery rate varied between 75% and 102% (Huerta-Fontela et al., 2011).
It is worth noting that AEDs and their metabolites/transformation products should be analyzed by different operation modes of LC-MS according to analysis strategies. For quantification of drugs, the most common mode operation is monitoring reaction multiple (MRM) of target analysis. The relevant standards in this mode are required to optimize compound-specific MS conditions including fragment-ion masses, ion-source voltages and collision energy (Acena et al., 2015). And environmental concentrations can be defined according to standard curve. As for the identification of unknown products, the operation modes of suspect and nontargeted screening will be recommended. The modes need to acquire mass spectrum information, retention time and isotope pattern by UPLC-high resolution mass spectrometry (UPLC-HRMS), then match the existing mass spectrum library such as MassBank, mzCloud, and so on, and finally define the compounds with the highest similarity of mass spectrum (Du et al., 2022). In previous study, twenty-two transformation products of CBZ have been found by nontarget analysis during chlorination and five of those were first reported (Han et al., 2018). However, specific structures of partial transformation products with two isomers such as chlorinated acridone (Cl-acridone) have not been identified limited to the current mass spectrometry analysis methods, which is also a challenge for nontarget analysis.
In the comprehensive analysis of the above literatures, the accurate analysis of AEDs can be achieved by the SPE preparation method with HLB as cartridge and methanol as elution solvent, and instrumental analytical methods with HPLC-MS/MS or UPLC-MS/MS. And the samples from sludge or sediment need be cleaned up by HLB cartridge to achieve more efficient recovery results. In addition, the choice of target, suspect and nontarget analysis should depend on the specific analysis purposes, which are essential to concentration definition and identification of metabolites/transformation products.

Environmental concentrations
The concentrations of AEDs in aquatic environments around the world from the published literature in recent years are listed in Table 3. The results showed that the environmental concentrations of AEDs in WWTP influents and effluents, surface water and drinking water ranged from ng L −1 to μg L −1 . Similarly, the metabolites/transformation products of AEDs in WWTP effluents ranged from ng L −1 to μg L −1 . Specific concentration characteristics were shown in the following sections.
Due to the increasing use of AEDs year by year and the fact that AEDs and their metabolites are often insufficiently removed while passing through WWTPs, the concentrations of partial AEDs like CBZ in the WWTPs effluents were relatively high and even exceeded the concentration in the influent in some cases. The most probable explanation is that carbamazepine glucuronides and other conjugated metabolites could transfer to the CBZ via enzymatic process during the activated sludge process (Gros et al., 2010); Another explanation is that previous activated sludge release the CBZ in their body (Vieno et al., 2007). However, the fact is that AEDs were discharged into the aquatic environments along with WWTP effluents (Gurke et al., 2015;Petrie et al., 2015;Verlicchi et al., 2012).

Surface water
Surface water mainly includes river and lake, which is supposed to important "sink" of AEDs in aquatic environments. CBZ is one of the most frequently detected drugs in surface water, although it is excreted in its unmetabolized form by only few percent of people (Ternes, 1998;.  mean concentration (419.7 ± 678.4 ng L −1 ) of CBZ in surface water was lower than that (654.3 ± 671.0 ng L −1 ) in WWTP effluents. The decrease in concentration was due to the dilution of rivers and degradation in natural environment. It only means CBZ is not detected at the level of instrumental analysis, but does not mean the disappearance of CBZ pollutants in nature.
More degradation products could form in water, which cause more serious environmental hazards. In addition, GAB (n.d.-1887 ng L −1 ) was found in the rivers of UK, which was lower than the concentrations (213-56810 ng L −1 ) of WWTP effluents. It worth noting that the concentration data of other AEDs such as lorazepam, diazepam, OXC, etc., are extremely scarce. Future researches similarly need to focus on analyzing their concentration levels.

Drinking water
Drinking water mainly includes ground water, drinking water sources and tap water in this study, which is supposed to be directly related to human exposure. The concentrations of CBZ were 900 ng L −1 in ground water (Sacher et al., 2001), 0.41-749 ng L −1 in drinking water sources (Huerta-Fontela et al., 2011;Kleywegt et al., 2011;Stackelberg et al., 2007;Sun et al., 2015; and 13-27 ng L −1 in tap water (de Oliveira et al., 2019). The mean concentration (203.7 ± 320.9 ng L −1 ) of CBZ in drinking water was lower than that (654.3 ± 671.0 and 419.7 ± 678.4 ng L −1 ) in WWTP effluents and surface water. In addition, diazepam, phenytoin and primidone were also detected in drinking water sources in Spain at concentrations of 12-200 ng L −1 (Huerta-Fontela et al., 2011). The decreasing trends on concentrations identified that WWTP effluents are the important source of AEDs indirectly. Although the dilution effects of river may reduce the concentration of AEDs, AEDs can still occur a migration in watershed scale due to their stability and further pollute the drinking water sources. Thus, the large-scale migration of AEDs needs to be further revealed in the future.

Transformation pathways by metabolizing in the human body
Generally, the parent chemicals of AEDs are also excreted from the human body with associated metabolites. CBZ was predominantly metabolized in the liver, and thirty-three metabolites of CBZ were identified from human and rat urine (Lertratanangkoon and Horning, 1982  first clinical studies of CBZ indicated that most (72%) of the consumed CBZ in a human was excreted via urine and few (28%) via feces (Bahlmann et al., 2014;Faigle and Feldmann, 1975). Commonly, AED metabolisms consists of phases I and II reactions (Figure 1), summarized by (Bahlmann et al., 2014). Phase I mainly lead to the formation of oxygenation metabolites. Hydroxylation and epoxidation were categorized as phase I reactions, and included three key metabolic pathways: first (A), CBZ may be transformed to CBZ-EP under the action of cytochrome P450 (Kerr et al., 1994). The CBZ-EP was further metabolized into 10,11-CBZ-diOH (Kitteringham et al., 1996). Both CBZ-EP and 10,11-CBZ-diOH can transform to 9-hydroxymethyl-10-carbamoylacridan (9-HMCA) under the function of ring contraction (Richter et al., 1978); the second metabolic route (B) involved the formation of intermediate of hydroxylated CBZ, such as 2-CBZ-OH, 3-CBZ-OH, then 2,3-dihydroxycarbamazepine (2,3-CBZ-diOH), and finally lead to the formation of CBZ-o-Quinone (Pearce et al., 2002); the third metabolic route (C) lead to the formation of trace 2-hydroxyiminostilbene, iminoquinone, 9-aldehyde-acridine, acridine and acridone (Furst and Uetrecht, 1993;Ju and Uetrecht, 1999). The phase II reactions were mainly that CBZ and CBZ-EP transformed to CBZ-N-glucuronides by carboxamide group (Bahlmann et al., 2014;Bauer et al., 1976;Maggs et al., 1997). The metabolic processes of AEDs are different in the human body due to their characteristics of chemical structures and properties. GAB cannot be metabolized in the body and was excreted in almost unchanged form via urine (Henning et al., 2018). Similar with CBZ, large parts of OXC were transformed to the main metabolite 10-CBZ-OH with antiepileptic effects, which was subsequently oxidized to 10,11-CBZ-diOH. Only 2-4% of OXC was excreted as the unchanged drug (Kaiser et al., 2014).
Furthermore, CBZ was also founded to be degraded by fungi. For example, Jelic et al. described the aerobic degradation of CBZ by the white-rot fungus Trametes versicolor (Jelic et al., 2012). Kang et al. identified the fungal metabolites of CBZ, in which two model fungi Cunninghamella elegans ATCC 9245 and Umbelopsis ramanniana R-56 were tested for metabolisms of CBZ. The results showed that both fungal strains produced CBZ-EP as a major metabolite and 3-CBZ-OH as one of the minor metabolites. The degradation of CBZ by C. elegans ATCC 9245 and U. ramanniana R-56 likely occurred because of mono-oxygenation reactions, which were similar to mammalian metabolisms (Kang et al., 2008). Besides, some reports indicated that CBZ can be degraded by other fungi and bacterium, such as Phanerochaete chrysosporium (Rodarte-Morales  (Kerr et al., 1994;Kitteringham et al., 1996;richter et al., 1978); Phase i (B) referred to (Pearce et al., 2002); Phase i (C) referred to (Furst and Uetrecht, 1993;Ju and Uetrecht, 1999); Phase ii referred to (Bahlmann et al., 2014;Bauer et al., 1976;maggs et al., 1997Zhang and Geissen, 2012), Aspergillus niger and Rhodococcus rhodochrous (Gauthier et al., 2010).
As for other AEDs, the studies were relatively less than CBZ and OXC. Diazepam will transform to a monohydroxylated product or an N-demethylated product. The demethylated product is a prescribed sedative, and supposed to be an active human metabolite of diazepam. Levetiracetam will transform to their corresponding carboxylic acids after primary amide hydrolysis (Helbling et al., 2010). Figure 3 showed the transformation pathways of CBZ by water treatment processes including advanced oxidation processes (AOPs) and chlorination disinfection processes. AOPs have proved to be an effective method for the transformation of many toxic and nonbiodegradable organic contaminants. Graphitic carbon nitride with plentiful nitrogen vacancies (NV-g-C 3 N 4 ) was used for constructing a photocatalytic system to degrade CBZ in the presence of sulfite under visible light. The photocatalytic system was found to be efficient for the transformation of CBZ, and a transformation pathway was proposed (Cao et al., 2019). First, CBZ transformed to the CBZ-EP because of the direct attack of the C = C on the central azepine ring by sulfite radicals (SO 3 ·− ). Then, the CBZ-EP was further transformed via the attacking of SO 3 ·− and the hydrolysis of the urea group on the central heterocyclic ring of CBZ, in which some intermediates were produced such as 9-carbaldehyde-acridinium, 10,11-CBZ-diOH, 1-(2-formylphenyl)urea, benzoic acid, succinic acid, terephthalic acid and formic acid.

Transformation pathways by water treatment processes
Another efficient photocatalytic heterostructural TiO 2 -based nanocomposite was synthesized using a simple hydrothermal method (Shahzad et al., 2018). The composite exhibited excellent photocatalytic activity for CBZ transformation and two different transformation pathways were proposed based on the detected intermediates. In path I, the hydroxyl radicals (HO · ) substitution in the CBZ molecule could generate some intermediates such as 10,11-CBZ-diOH and 2,2′-(carbamoylazanediyl)dibenzoic acid. Then, the intermediates were followed by ring cleavage to produce 2-hydroxybenzoic acid and 2-aminobenzoic acid, which were further transformed to aniline and benzoic acid and ultimately degraded into CO 2 and H 2 O. In path II, the CBZ-EP were produced by the · O, leading to the generation of acridine. Further, the HO · attacked acridine and produced formaldehyde-acridine. Besides, Tran et al. reported a sonochemical transformation process of CBZ in an aqueous solution using ultrasound (Tran et al., 2013). And the removal rate of CBZ can reach 90.1%. The oxidant species were HO · produced in the sonolysis reactor, and CBZ was primarily transformed into anthranilic acid and acridine. The proposed transformation mechanisms of CBZ were similar with the report by Shahzad et al. (Shahzad et al., 2018). Furthermore, other AOPs such as O 3 (Mcdowell et al., 2005), UV/TiO 2 (Doll and Frimmel, 2005), ultrasonic/Fe 0 /H 2 O 2 (Ghauch et al., 2011), BiOCl/AgCl (Meribout et al., 2016), UV/chlorine (Zhou et al., 2016), UV/H 2 O 2 (Vogna et al., 2004) and activated persulfate (Deng et al., 2013) have proven to be efficient processes in the transformation of CBZ. All these AOPs mentioned above are generally characterized by the generation of highly reactive radicals, which can rapidly oxidize CBZ and produce similar intermediates.
In terms of three transformation pathways by metabolizing in human body, microorganisms and water treatment processes, the transformation of AEDs was always in the direction of oxidation. Specifically, hydroxylated and ketonizated products was conductive objective. And there are a lot similarities in three transformation pathways. As an example of CBZ, CBZ-EP and 10,11-CBZ-diOH act as the intermediates, and the final products include acridine and acridone in all transformation pathways. It indicates that there can be existing the same reaction mechanisms in different transformation environments while the species were different. Meanwhile, the difference of transformation pathways was also observed. 9-HMCA, iminoquinone and CBZ-N-glucuronide were only found in the transformation pathways by metabolizing in human body, which can be supposed to be the markers in the identification of the source. Analogously, BaQD and BaQM were only found in the transformation processes of microorganisms. And more products of small molecules such as 1-(2-formylphenyl)urea, 2-hydroxybenzoic acid and 2-aminobenzoic acid occurred in the water treatment processes. Thus, these distinctive chemicals can be the evidence for identifying the source of AEDs and their metabolites/transformation products.

Fate of AEDs in aquatic environments
The transportation and transformation processes of pharmaceuticals are very complex in aquatic environments. Considering most AEDs were refractory to degradation in the water treatment processes, here we mainly discuss the natural processes of the parent AEDs in aquatic environment as some transformation products may occurred in the above processes. Usually, partitioning (sorption and desorption) and degradation (abiotic and biotic degradation) determine the fate of pharmaceuticals after they enter aquatic environments. Partitioning is an important process that affects the fate of pharmaceuticals in aquatic environments. The sorption processes are mainly due to hydrophobic interactions and other binding processes such as cation exchange, cation bridging, surface complexation, and hydrogen bonding. The sorption processes are not only relevant to the solubility or octanol-water partition coefficients (Log K ow ), it is also affected by the specific water quality including concentration and nature of suspended solid particles in the water phase, pH and organic matter content of the aquatic environment (Ying et al., 2013). The abiotic degradation processes mainly include photolysis and hydrolysis. The photolysis processes can be divided into direct photolysis and indirect photolysis, in which the direct photolysis of pharmaceuticals is caused by direct absorption of solar light and the indirect photolysis is caused by strong oxidant species produced by the natural photosensitizers such as nitrate and humic acid (Andreozzi et al., 2003). Hydrolysis can be another major route for pharmaceutical degradation in aquatic environments, but some exceptions exist in which the pharmaceuticals such as steroids and acidic drug cannot undergo hydrolysis in water (Ying et al., 2013). The biotic degradation involves different levels of organisms such as bacteria, fungi, and algae. Chemical structures of AEDs and environmental conditions are the key factors for biodegradation of a pharmaceutical in surface water (Ying et al., 2013).
Normally, the weaker the hydrophobicity interaction of pharmaceuticals, the less their Log K ow and the greater difficultly they have been adsorbed onto sediment. CBZ, with a predicted octanol-water partition coefficients (Log P) value of 2.77, has low hydrophobicity and is the most abundant pharmaceutical soluble in water phase. The results of Yuan et al. similarly showed that the concentrations of CBZ in water phase were relatively higher than those in sediment and suspended particulate matter (Yuan et al., 2019). CBZ, as a kind of AEDs for oral intake, are considered resistant to hydrolysis. Thus, photolysis seems to be the main pathway in abiotic degradation (Andreozzi et al., 2003). Several studies also suggested that the hydrolysis of CBZ can be neglected due to its resistance against strong bases and acids (Tixier et al., 2003;Yuan et al., 2019). In addition, CBZ is refractory to biodegradation in aquatic environments. Thus, photolysis processes may be the most important degradation pathway for CBZ in aquatic environments. Previous studies revealed that CBZ can be degraded by direct or indirect photolysis, and several constituents in the natural water such as chloride and dissolved organic matter enhanced the photodegradation of CBZ (Chiron et al., 2006;Lam and Mabury, 2005;Matamoros et al., 2009). Moreover, more evidence indicated that direct photolysis can be the main degradation pathway (Yuan et al., 2019). De Laurentiis et al. reported the photochemical fate of CBZ in surface water (De Laurentiis et al., 2012). The results indicated that the direct photolysis and HO · reactions were the main CBZ photodegradation pathways. The main quantified photodegradation intermediates for direct photolysis and HO · reaction were acridine (3-3.5%) and 10,11-CBZ-diOH (4-8.5%). The other important photo-intermediates were similarly found. For example, aromatic-ring-CBZ-diOH was hydroxylated twice directly by CBZ. 1,1-bis(2-formylphenyl) urea was formed by CBZ-EP and further formed BQD, which was similar as the products of chlorination and chloramination (De Laurentiis et al., 2012;Han et al., 2018). And OXC also showed the same photolysis intermediates and products as CBZ including 1,1-bis(2-formylphenyl) urea and BQD (Hernández-Tenorio, . The fate of diazepam in surface water is similar with that of CBZ but with some differences. It was similarly noted that biodegradation and mineralization were not important environmental fates for diazepam in sediment/water systems. Diazepam is supposed to be resistant to hydrolysis because it does not have hydrolyzable bonds in the molecular structure. But diazepam can slowly partition on the sediment in sediment/water systems, and aquatic photodegradation was estimated to be an important environmental fate process for diazepam in the top layer of surface water. Thus, most diazepam was expected to slowly adsorb to the undissolved solids in the water column or it may adsorb directly to the sediment, while a small fraction of diazepam can be photodegraded rapidly in the superficial layers after it entered the receiving waters (West and Rowland, 2012). Their results also indicated that the photolysis processes of diazepam and lorazepam under simulated sunlight in water cannot be neglected (Calisto et al., 2011;West and Rowland, 2012).
The environmental risks of AEDs and their transformation products were evaluated according to the collected data. The calculation methods of risk quotient (RQ), toxicity data and results were described in Supporting Information (SI) Text S1, Tables S1-S3. In general, RQ > 1 indicates high ecological risk, 0.1 ≤ RQ ≤ 1 indicates medium ecological risk, and RQ < 0.1 indicates low ecological risk (Hernando et al., 2006). RQ calculated by chronic toxicity (RQ C ) was discussed in this study based on the conservative consideration. In terms of WWTP effluents, RQ C values of five AEDs and their transformation products including lamotrigine (22.4), CBZ (6.8), OXC (4.3), 10-CBZ-OH (1.8) and CBZ-EP (1.0002) were more than 1.0, which indicates a high ecological risk. And primidone has RQ C value between 0.1 and 1.0, which indicates a medium ecological risk. As for surface water, ground water and drinking water sources, two AEDs including phenytoin (26.7) and CBZ (3.0) have RQ C values exceeding 1.0, which indicated a high ecological risk. Previous studies also revealed that diazepine and CBZ can induce a medium and high risk from effluents (Kondor et al., 2022;Verlicchi et al., 2012).
Although the environmental risks induced by AEDs may be a consensus issues, their transformation products similarly can cause serious environmental hazards, which needs to attract more attention.
In a word, as an example of CBZ and diazepam, these AEDs were resistant to hydrolysis. And they dominated the absolute concentration distribution in water phase than in sediment phase. Furthermore, more evidence indicated that photolysis was an important process for tested pharmaceuticals in the photic zone of receiving waters, and the humic acids and fulvic acids can affect the photolysis processes. But it worth noting that photodegradation rate of AEDs have largely difference such as the half-time life of CBZ (110 days) and diazepam (7 days) (Calisto et al., 2011;De Laurentiis et al., 2012). Even so, CBZ and diazepam can also exist in the aquatic environment with the original face and photodegradation products for a long term. It also gives us a remind to monitor the real form of these AEDs in aquatic environments.

Conclusions and perspectives
Traditional WWTPs are not completely capable of removing AEDs or their metabolites/transformation products during treatment processes, and the treated effluents discharged into receiving water bodies may still contain a certain concentration of AED residues. AEDs and their metabolites/transformation products are widely found in aquatic environments such as surface water, groundwater and even drinking water with concentrations between ng L −1 and μg L −1 . The accurate target analysis of AEDs can be achieved by the SPE preparation method with HLB as cartridge and methanol as elution solvent, and instrumental analytical methods with HPLC-MS/ MS or UPLC-MS/MS. AEDs and their metabolites/transformation products indeed influence aquatic environments and organisms to a certain extent. Partial AEDs including CBZ, GAB, OXC and diazepam can commonly cause the damage of antioxidant capacity, which were paid more attention. In addition, considering the endocrine disrupting effect of carbamazepine on fish, the reproductive effects of AEDs and their metabolites/transformation products on aquatic organisms also need to be concerned in the future. However, toxicity data of metabolites/transformation products of AEDs was extremely deficient, which was mainly because of incomplete transformation pathways and absent standards. Therefore, more toxicity data including AEDs and their metabolites/transformation products need to be acquired to evaluate their environmental hazards.
In aspect of metabolisms/transformation mechanisms in human body, microorganisms and water treatment processes, the transformation of AEDs was always in the direction of oxidation, and hydroxylated and ketonizated products were conductive objectives as the difference of transformation pathways was also observed. These differential products can be supposed to be the indicators in the identification of the source. However, there may be still many products unfound limited to the analytical methods and standards. Thus, the detected concentration cannot entirely present the actual influence that this compound caused. And in all objective AEDs, CBZ have more comprehensive descriptions on transformation pathways than other AEDs like OXC, diazepam and levetiracetam. Products and transformation pathways of other AEDs need to be further explored.
The fate of AEDs in natural water mainly depends on their inherent properties (chemical structure, Log K ow , etc.). Because of their low Log K ow and hydrophobicity, AEDs were supposed to dominate the absolute concentration distribution in water phase than in sediment phase. Furthermore, photolysis was a main process for AEDs in the photic zone of receiving waters. But photodegradation rates of AEDs were still low, which mean that they can also exist in aquatic environments with the original face and photodegradation products for a long term. It also gives us a remind to monitor the real form of these AEDs in aquatic environments. Furthermore, RQ C values of five AEDs and their transformation products including lamotrigine, CBZ, OXC, 10-CBZ-OH and CBZ-EP in WWTP effluents can induce a high ecological risk, which should receive more attention.
Another clear demand exists for the development of advanced wastewater treatment technologies to more efficiently and lower costly remove/degrade AEDs and their metabolites/transformation products as AOP can be suitable to transform the AEDs to the small molecular compounds with low toxicity. Partial studies also showed that constructed wetland and photocatalysis presented good removal rates of AEDs (Daniel Cardoso-Vera et al., 2021;Dordio et al., 2010). It worth noting that the complexion of aquatic environment is required to be considered. The existence of interfering substances (bromide, iodide, etc.) will lead to the formation of higher toxic compounds (Heeb et al., 2014;Hua et al., 2006). Thus, these brominated or iodinated products of AEDs need to be also noted.
In the future, more studies should focus on the transformation mechanisms, which is necessary to monitor products of AEDs. Nontarget analysis technologies by means of UPLC-HRMS or comprehensive two-dimensional gas chromatography-time of flight-mass spectrometry (GC × GC-TOF MS) are powerful tools to identify unknown products and intermediates (Du et al., 2022). Several studies have revealed formation pathways of polycyclic aromatic hydrocarbons derivatives and halophenylacetamides during chlorination using nontarget analysis (Hu et al., 2022;Liu et al., 2020). It showed that nontarget analysis has significant advantages in identification of unknown products. Furthermore, development of structure prediction software (CSI:FingerID/SIRIUS 4, MSNovelist, etc.) based on machine learning algorithm will also provide efficient supports on spectral analysis of unknown products (Duhrkop et al., 2019;Stravs et al., 2022), which enable researchers to obtain the corresponding standards more accurately. Thus, it is recommended to combine mechanisms analysis with nontarget analysis to identify the metabolites/transformation products.