Glycosylation was observed for diclofenac, sulfamethoxazole and di-n-butyl phthalate in A. thaliana, triclosan, naproxen, diclofenac, ibuprofen and gemfibrozil in carrot cell cultures,bisphenol A and carbamazepine in lettuce,and TBBPA in pumpkin seedlings.Acetaminophen and chlortetracycline were conjugated with glutathione in cucumber seedlings and maize seedlings, respectively.Conjugation with other biomolecules, such as saccharides, malonic acid, and sulfate, was also occasionally reported for some CECs, such as triclosan in carrot cells and diclofenac in A. thaliana cells.Other than conjugation, methylation is also a type of phase II metabolism and has been reported for TBBPA in pumpkin seedlings. The investigation of phase III metabolism of CECs in plants is relatively limited, as quantitative evaluation of phase III products would require the use of isotope labeling to account for the nonextractable or bound residues, although phase III is expected to be dominant in determining the final destination of xenobiotics in plants. For instance, nearly all 14C- labeled naproxen, diclofenac, bisphenol A and nonylphenol were found in nonextractable bound residues in lettuce and collards.Numerous studies have documented the occurrence of CECs in aquatic environments in many countries and regions. CECs are introduced into aquatic environments via discharge of TWW from WWTPs, agricultural activity, landfill leachates, and surface runoff . Concentrations of CECs ranges from ng L-1 to μg L-1 in impacted surface water and from μg kg-1 to mg kg-1 in the sediment. The concentration of CECs in surface water is largely influenced by the population density, environmental conditions and terrigenous supply, displaying temporal and spatial variations. Studies conducted in the U.S. in recent decades have shown the occurrence of hundreds of CECs in various watersheds, with the maximum concentration at 35 μg L-1 for sucralose in water. A comprehensive review of the occurrence of CECs in Latin America, including studies performed in 11 different countries between 1999 and 2018,vertical planters for vegetables has shown the common detection of 17β-estradiol, bisphenol A and estrone.
The highest concentration of CECs detected reached 1100 μg L-1 for clindamycin in Costa Rica.Many survey studies have also been reported for countries in Europe, such as the Sava River in Slovenian and Croatian,rivers receiving TWW in Ireland,and impacted rivers and lakes inSweden.An EU-wide survey of CECs in European river waters indicated that benzotriazole, caffeine, carbamazepine, tolyltriazole, and nonylphenoxy acetic acid were among the most frequently detected and/or at the highest concentrations.In addition, levels of CECs in some developing countries may be higher than those in developed countries, likely due to less rigorous treatment at WWTPs. For example, naproxen as high as 140 μg L-1 was reported in a study originated in India and up to 167 μg L-1 for lamivudine in Kenya.Research focusing on the occurrence of CECs in sediment is less prevalent and typically involves fewer CECs. Furthermore, the concentration of CECs in sediments usually exhibits less seasonal variations, which suggests that sediment samples may serve as a more stable marker for CEC monitoring in aquatic environments.High detection frequencies and concentrations of estrone and 17β-estradiol have been reported in sediments from the mouth of the Manokin River in the U.S., with the highest concentration at 58.4 μg kg-1 and 11.5 μg kg-1 , respectively.In the Southern California Bight, triclosan, 4-nonylphenol and bis have been detected in all sediments at median concentrations of 5.1 , 30 , and 121 μg kg-1 , respectively.The highest total concentration of antibiotics in sediments from the intertidal zones of the Yellow River Delta, China was measured to be 178.77 μg kg-1 . 124 Studies conducted in African countries such as Morocco showed even greater CEC concentrations in the sediment . Interestingly, several surveys have shown the presence of some hydrophilic CECs in sediments, such as acetaminophen and caffeine, which were previously thought to have limited sorption to solids due to their low hydrophobicity.Aquatic organisms living in impacted aquatic environments have been sampled for detection of CECs. CECs have been found in the tissues of fish, mussels and oysters collected from the impacted water systems in the U.S., with the maximum concentration detected at 3000 ng g-1 for 4-nonylphenol in mussels, suggesting potential bioaccumulation of CECs in aquatic organisms.
Diazepam was detected in all collected flatfish liver samples in Southern California but was infrequently detected in sediments, highlighting the bio-magnification potential of certain CECs.The accumulation of PBDEs in fish livers was comparable to that of legacy organochlorines.Water snakes and small common carps living in an e-waste contaminated water pond were reported to accumulate plasticizers and organophosphorus flame retardants in their tissues.Therefore, aquatic organisms may be exposed to low levels of CEC mixtures in the environment, and bio-accumulation in aquatic organisms is possible for some CECs. As CECs are continuously introduced into aquatic environments via various pathways, they may be considered as pseudo-persistent contaminants, causing long-term, mixed, low-dose exposure to aquatic organisms. A comprehensive review of the ecotoxicity of human pharmaceuticals concluded that for all human medicines tested, acute effects to aquatic organisms were unlikely, unless spill incidents occurred, due to their trace level occurrence in aquatic environment.The chronic lowest observed effect concentrations of most tested pharmaceuticals in standard laboratory organisms are about two orders of magnitude higher than what detected in the effluents.However, recent studies showed that some CECs may exert unintended adverse effects to organisms, such as endocrine disruption and developmental toxicity, at environmentally relevant levels.The investigation on toxic effects of CECs on aquatic organisms includes two main types of exposure: direct exposure to the real environment, such as TWW and impacted water, and exposure to water spiked with CECs under controlled conditions. Chronic effects, including sublethal effects, have been often observed in aquatic organisms exposed to affected water bodies or TWW. For example, fathead minnows and freshwater mussels were caged for 4 weeks upstream and downstream of the discharge from WWTPs, and were found to develop multiple biomarker responses, such as oxidative stress, enzyme induction, shifts in gene expression and alteration of immune functions.
The growth and yield of green algae and reproduction of daphnia were inhibited by TWW and exhibited dose-response effects. Juvenile rainbow trout exposed to TWW showed significantly different plasma cortisol and glucose response to the secondary stressor.However, it is often difficult to interpret the impact of CECs in this type of experiments, as various other stressors, such as water temperature and bacteria in the real environment, may also induce such biomarker responses.The low concentrations of CECs in TWW also could not explain the sublethal effects observed on algae and daphnia.In addition, interactions of compounds in CEC mixtures should be further considered. The exposure of aquatic organisms to artificially spiked CECs, on the other hand, provides comparable toxicological data under controlled conditions. Most research has been devoted to the toxic effects of CECs at the individual level, while in realistic situations, CECs are always present as a mixture. Exposure to CECs at environmental relevant levels cause multiple adverse effects. For example, marine mussels exposed to atorvastatin at around 1.2 μg L-1 exhibited key fatty acid metabolism disruption and suppression of xenobiotics efflux through P-glycoprotein and membrane diffusion.Gemfibrozil was shown to reduce plasma androgens in goldfish after exposure to 1.5 μg L -1 for 4 and 14 days;while the concentration of gemfibrozil in WWTP effluent was found to be in the range of 10-3830 ng L-1 . The adverse effects of CECs have also been shown at the population level. For example, a 7-year, whole lake experiment conducted in northwestern Ontario, Canada, showed that chronic exposure of fathead minnow to 5-6 ng L-1 of the synthetic estrogen, 17 α- ethynylestradiol,vertical farming technology led to the near extinction of this species.Aquatic invertebrates have been widely adopted to derive acute toxicity end-points, e.g., LC50 values, for target CECs. The acute toxicity of CECs varied greatly, even for compounds belonging to the same chemical class and displayed species-specific effects. For example, the EC50 and LC50 values varied largely among the 12 tested polychlorinated diphenyl ethers for S. obliquus, D. magna, and D. rerio, respectively.Exposure to 17α-ethinylestradiol, acetylsalicylic acid, and bisphenol A significantly affected the embryonic development of sea urchins, with different LC50 values for Mysidopsis juniae and Artemia sp. The concentrations that induced 50% growth inhibition in algae of metolachlor, erythromycin, and triclosan also showed multiple-fold differences between freshwater and marine algae, reflecting the species-specific sensitivity.Mixed exposure of silver nanoparticles, polystyrene nanoplastics and 5-fluorouracil displayed interaction toxicity to marine mussels, with exponentially increased oxidative damage compared to individual contaminants,highlighting the importance to consider chemical interactions when investigating the toxic effects of CECs in the real environment. Some government agencies in the U.S., such as EPA and California State Water Resources Control Board, have tried to put some regulations to control CECs in aquatic environments. For example, the Science Advisory Panel for CECs in California’s aquatic ecosystems has developed strategies to identify the monitoring trigger levels of CECs in aquatic environments based on their lowest effect values available from established databases, such as the Computational Toxicology database and the NORMAN database , for aquatic organisms.However, for the TPs of CECs, such data are usually not experimentally available. Due to the huge and continually increasing number of CECs in aquatic environments, it is unrealistic to examine the toxicity effects of all CECs. Several studies have attempted to develop a prioritization process to select CECs that require the most attention for aquatic organisms based on their monitoring data, production volume, persistence and prevalence in the environment, bio-accumulation potential, and biological effects.
Several modeling tools, like machine learning and quantitative structure-activity relationships , have also been developed to predict the bio-accumulation, bio-transformation and toxicological effects of CECs.For example, Sequence Alignment to Predict Across Species Susceptibility was adopted by the Science Advisory Panel for CECs in California’s aquatic ecosystems to predict the behaviors of CECs across species without available toxicological data from the existing database.The incorporation of such tools is of vital importance to improve risk assessment of CECs due to the limited experimental resources. Studies have often revealed that TPs of CECs occur simultaneously in the tissues of aquatic organisms with their parent compounds, sometimes at even higher concentrations. For instance, metabolites of organophosphorus flame retardants were found in the same order of magnitude as their parent compounds in water snake and small common carps collected from an e-waste-affected site. Norsertraline, the demethylated TP of sertraline, was found to be bio-accumulated at a greater degree than sertraline in the liver of rudd collected from the TWW-impacted Niagara River.Nordiazepam, the demethylated TP of diazepam, was also frequently detected in aquatic organisms along with diazepam.127 Therefore, TPs of CECs in aquatic organisms could originate from two sources – uptake from the ambient environment, and transformation taking place in the organism upon the uptake of the parent compound. Research focusing on the biotransformation of CECs in aquatic organisms, including aquatic plants such as algae, invertebrates such as daphnia, and vertebrates such as fish, is limited. However, the identification of CEC metabolites in aquatic organisms is crucial for evaluating the ecological risks of CECs. Prevalent phase I and phase II enzyme activities were frequently induced in aquatic organisms after CEC exposure, such as cytochrome P450 enzymes and glutathione transferases.Although some common metabolites might be expected, the pattern of CEC metabolism could also vary between different species. For example, three phase I metabolites and 10 phase II metabolites were identified in marine mussels exposed to diclofenac,while 7 phase I metabolites and 3 phase II metabolites were found in H. Azteca and G. pulex. Significant differences in biotransformation rates were observed for different species or between opposite sexes of fishes exposed to CECs.157 Certain aquatic species, such as glass eels, displayed low metabolic activity, with few metabolites detected after CEC exposure,155 while the absence of bio-magnification effects of PFRs in water snakes was attributed to the active bio-transformation. Bio-transformation of CECs in algae shared some similar pathways as that in terrestrial plants, as in the case of hydroxylation, demethylation and glycosylation of bisphenols.Methylation and demethylation are common transformation pathways for chemicals in the environment, especially for compounds with -OCH3, -NCH3-, -SCH3, and/or the corresponding -OH, -NH-, and -SH groups in their chemical structures.