Biological and Environmental Impact of Pharmaceuticals on Marine Fishes: A Review

Abstract

Pharmaceuticals are recognized as a serious threat to aquatic ecosystems due to their persistence or pseudo-persistence and their biological activity. Their increased consumption in human and animal medicine has led to a continuous discharge of such biologically active molecules in aquatic environments. Marine ecosystems have been poorly investigated, even though recent studies have confirmed that these emerging contaminants occur widely in these ecosystems. Due to their interaction with specific biochemical and physiological pathways in target organisms, pharmaceuticals can cause alterations in several marine species during their entire life cycle. In particular, marine fishes have shown the ability to bioaccumulate these compounds in their body, and they may be used as potential bioindicators of pharmaceutical contamination in seawater. The objective of this review was to provide a comprehensive overview of the current understanding of the sources and occurrence of pharmaceuticals in marine environments, illustrating the adverse biological effects of important classes of these compounds on marine fishes.

1. Introduction

The advancement of medical science and the development of efficacious pharmaceuticals have resulted in a reduction of the impacts of several pathologies, thereby improving life expectancy, prevention of disease, and health quality [1]. In consequence, the worldwide pharmaceutical industry has exhibited substantial expansion, both in the volume of sales and the number of synthesized active pharmaceutical substances, due to ongoing development as well as the aging of the global population [2]. Pharmaceuticals have been identified as having a beneficial role in human society. However, the widespread consumption of these emerging contaminants has become a significant threat to aquatic ecosystems due to their biological activity, their frequency of detection in the environment, and the potential negative effects on non-target species [3,4]. Some pharmaceuticals are defined as having pseudo-persistence, while other pharmaceuticals remain persistent to environmental degradation in their original form or as metabolites and are called “Environmentally Persistent Pharmacological Products” (EPP) [5].

Approximately 4000 pharmaceuticals are utilized globally for the treatment of human and animal ailments, as well as the promotion of livestock’s growth [6,7,8,9,10,11,12]. A considerable portion of pharmaceuticals present in the human or animal body is excreted via urination or defecation, or in their native form or as products of degradation following metabolic reactions [13,14]. Secondary sources of pharmaceuticals in the environment context comprise land application of sewage sludge, manufacturing facilities, and agricultural, aquacultural, and veterinary practices [15]. The high solubility in water and the presence of polar functional groups in their chemical structures facilitate the discharge of pharmaceuticals and their metabolites into surface water, despite their inefficient removal in wastewater treatment plants (WWTPs) [16]. Recently, it has been reported that the outbreak of the coronavirus disease-2019 (COVID-19) caused by the severe acute respiratory coronavirus 2 (SARS-CoV-2) represents a serious concern worldwide due to the massive use of COVID-19-related pharmaceuticals and their discharge into aquatic environments [17,18]. Since the onset of the COVID-19 pandemic, numerous therapeutic drugs have been used for the treatment of the increased number of COVID-19 patients [19,20]. A significant proportion of these pharmaceuticals are excreted from the human body and released into wastewater [21,22]. Moreover, the removal of these pharmaceuticals is not enhanced by conventional wastewater treatment plants, which are designed for removing specific compounds [23,24]. Consequently, the presence of COVID-19-related pharmaceuticals represents a significant concern for aquatic environments due to their biological characteristics and persistence [22,25,26].

The prevalence of pharmaceuticals in water bodies has been extensively investigated in both freshwater and groundwater [27]. However, their presence in marine ecosystems was only discovered in the early 2000s, and their effects have been poorly investigated due to the complexity of the matrices used to evaluate them and of their dilution effects in seawater [28,29].

Nevertheless, there is now a substantial corpus of research articles that have focused on the occurrence of pharmaceuticals in marine environments, which serve as the primary recipient of continental contamination. In particular, some published review articles have investigated specific pharmaceuticals, such as the persistence of diclofenac [30] or antibiotics [28], in marine compartments and the contamination of selected areas, including the Mediterranean Sea and the Arctic environment [31,32]. In order to enhance the detection, assessment, comprehension, and prevention of the pharmaceuticals’ adverse effects on the environment, a number of international agreements have been established. However, despite the establishment of these agreements, only a limited number of pharmaceuticals have been identified as contaminants of marine resources. The OSPAR Convention has established guidelines and recommendations for the monitoring of chemicals in marine waters of the Northeast Atlantic Ocean, including twenty-two pharmaceuticals and five hormones on the list of substances of potential concern [3]. The European Union has incorporated a number of pharmaceuticals into the Watch List of the Water Framework Directive (WFD) 2000/60/EC, which pertains to both freshwater and transitional areas [33]. To date, eight pharmaceuticals have been included in the list: three hormones (estrone (E1), 17β-estradiol (E2), and 17α-estradiol (EE2)) and five antibiotics (amoxicillin, ciprofloxacin, and the macrolides azithromycin, clarithromycin, and erythromycin) are potential candidates for the “Priority Substances List” [3,33,34]. Furthermore, it is imperative that other pertinent legislation be amended in accordance with the Marine Strategy Framework Directive (MSFD) 2008/56/EC. Pharmaceuticals are defined as biologically active compounds that can specifically interact with the physiological pathways of the targeted organisms. However, they can also cause additional impacts in marine environments, interfering with the health of non-targeted organisms already affected by climate change, eutrophication, and overfishing [1,31,35]. It has been reported that some therapeutic compounds can act at sublethal concentrations and alter biochemical and cellular responses, thereby influencing vital functions such as reproduction, metabolism, growth, feeding, immunity, and locomotion [3]. In the literature, there are numerous studies demonstrating the importance of aquatic organisms as bioindicators for evaluating the effects of different stressful conditions [36,37,38,39,40,41,42,43,44,45,46,47]. Among marine organisms, fish may be considered potential bioindicators of pharmaceutical pollution in the marine environment due to their life history and their susceptibility to drug uptake and its effects [48]. It can be posited that the aforementioned organisms may also be subject to the effects of residual pharmaceuticals, as they share a multitude of drug targets with humans [49]. Moreover, fish exhibit a considerable degree of morphological, physiological, and behavioral diversity, despite the paucity of research investigating the specific uptake, bioaccumulation, and tissue partitioning of pharmaceuticals in these species. A number of studies have reported the presence of pharmaceuticals in fish collected from different surface waters [50,51,52,53,54,55,56,57,58], and many studies have focused on the effects of antidepressants, antibiotics, and antihistamines in wild fish [52,59,60].

The objective of this study was to conduct a comprehensive analysis of the occurrence of some of the most prevalent classes of pharmaceuticals in marine environments, emphasizing their detrimental effects on non-target organisms. In the initial section of the article, we concentrate on the sources and persistence of pharmaceuticals in seawater. In the subsequent section of the review, we analyze the biological consequences of these compounds in marine fishes, identifying knowledge gaps that could be addressed in future studies.

2. Pharmaceuticals in the Marine Environment

2.1. Concentrations and Origin of Pharmaceuticals in Seawater

In general, the concentration of pharmaceuticals in seawater ranges from a few ng/L to hundreds of µg/L, with higher levels detected in freshwater than in marine ecosystems, depending on the physicochemical properties of the active ingredients and environmental factors [1] (Table 1, Table 2, Table 3, Table 4, Table 5 and Table 6).

Pharmaceuticals are excreted in their native form or as metabolites after use and enter the marine environment through different pathways: (i) human domestic use, (ii) hospital use, (iii) veterinary use through aquaculture or from the terrestrial environment (e.g., livestock production or care of pets), and (iv) industrial and commercial activities based on pharmaceuticals [29,61]. The most important pathway is represented by WWTPs, particularly domestic sewage treatment plants. The removal of organic matter and suspended solids is accomplished through the application of both physical and biological remediation techniques, which are used throughout the primary and secondary treatment processes [62]. Some studies have reported that physical remediation is based on coagulation, sedimentation, and flocculation, while biological degradation occurs with aerobic microorganisms that consume organic matter. Although these treatments are affected by several variables, such as the physicochemical characteristics of the pharmaceuticals, the characteristics of the sludge, and the technical characteristics of the reactors, the majority of pharmaceuticals are partially removed by primary and secondary processes that characterize conventional wastewater treatment plants [63,64,65]. In addition, changes in the volume of influent wastewater during intense rainfall events and/or seasonal increases in the population in tourist areas can significantly affect the removal capacity of WWPTs [66,67].

Wastewater from healthcare and pharmaceutical production is considered to be another important contributor of pharmaceuticals to marine and coastal environments. Hospital activities generate significant amounts of contaminated wastewater, depending on the number of patients or medical specialties [61]. A study by Oliveira et al. (2017) [68] showed that many pharmaceuticals were detected in hospital and healthcare effluents at concentrations below 10 µg/L, with high values for the most common agents, such as paracetamol (1368 µg/L) and ciprofloxacin (125 µg/L).

Large amounts of pharmaceuticals are also released from veterinary use in both aquaculture and land-based animal husbandry or livestock production [29,61]. Common veterinary pharmaceuticals used in aquaculture include antibiotics, analgesics, antiparasitics, and some generic human drugs used prophylactically [69,70,71].

Other important pathways for pharmaceuticals entering marine compartments are untreated wastewater discharges from land-based agricultural activities and livestock production, which include fecal excretion of grazing land metabolites and fertilizers [72,73,74]. Contamination of groundwater due to the presence of pharmaceuticals can also be linked to many other sources such as sewage contamination (via septic tanks or sewage leakage), contaminated leaks from landfills (i.e., animal carcasses and pharmaceutical waste products), and the use of fertilizers and grey water for irrigation [75].

Table 1. Concentrations of non-steroidal anti-inflammatory drugs in marine environments.

Non-Steroidal Anti-Inflammatory Drug	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Mefenamic acid	Mahdia Coast, Tunisia	Nd–0.6	[76]
	Eastern Mediterranean Sea, Greece	<0.2–11	[77]
	Gulf of Cadiz, Southwestern Spain	Nd–4.5	[78]
Phenylbutazone	Mahdia Coast, Tunisia	Nd–2	[76]
Phenazone	Gulf of Cadiz, Southwestern Spain	Nd–309.8	[78]
	Baltic Sea, Germany	5.9	[79]
	Aegean and Dardanelles, Greece and Turkey	2	[79]
Indomethacin	Gulf of Cadiz, Southwestern Spain	Nd–4.5	[78]
Tramadol	Eastern Mediterranean Sea, Greece	<0.1–1	[77]
Nimesulide	Portuguese seawaters	Nd–7.3	[80]
Codeine	Southwestern Taiwan	Nd–63.6	[81]
	Mediterranean coastal lagoon, Spain	1.8	[82]
Oxycodone	Mediterranean coastal lagoon, Spain	6.8	[82]
Acetylsalicylic acid	Portuguese seawaters	Nd–534	[80]
Acetaminophen	Gran Canaria Island, Spain	Nd–297	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–41.5	[78]
	Baltic Sea, Germany	48	[79]
	Aegean and Dardanelles, Greece and Turkey	2893	[79]
	Eastern Mediterranean Sea, Greece	<41	[77]
	Mediterranean Sea, Israel	12	[79]
	Seawater from Victoria BC, Canada	44.7	[84]
	San Francisco Bay, USA	85	[79]
	Southern California Bight, USA	Nd–11	[85]
	Portuguese seawaters	51–584	[80]
	Southwestern Taiwan	2.6–16.7	[81]
	Korean seawater	Nd–48	[81]
	Red Sea, Saudi Arabian coastal waters	2363	[86]
	Santos Bay, Brazil	Nd–34.6	[87]
Ketoprofen	Gran Canaria Island, Spain	Nd–106	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–2.6	[78]
	Southern Baltic Sea, Polish coastal zone	Nd–72.7	[88]
	Mahdia Coast, Tunisia	Nd–76	[76]
	Portuguese seawaters	10–90	[80]
	Southwestern Taiwan	Nd–23.3	[81]
	Northern Taiwan seawater	<1.7–6.59	[89]
Diclofenac	Gran Canaria Island, Spain	Nd–344	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–319	[78]
	Baltic Sea, Germany	9.2	[79]
	Mahdia Coast, Tunisia	Nd–23	[76]
	Eastern Mediterranean Sea, Greece	<1.4–16	[77]
	Mediterranean Sea, Israel	6.1	[79]
	Aegean and Dardanelles, Greece and Turkey	9.7	[79]
	Southern California Bight, USA	Nd–0.6	[85]
	Portuguese seawaters	Nd–241	[80]
	Santos Bay, Brazil	Nd–19.4	[87]
	Red Sea, Saudi Arabian coastal waters	14,020	[86]
	Seawater from Singapore	<2–12	[90]
	Marina Bay, Singapore	4–38	[91]
	Seawater from Northern Taiwan	<2.5–53.6	[89]
Naproxen	Eastern Mediterranean Sea, Greece	<0.01–0.8	[77]
	Gulf of Cadiz, South wstern Spain	Nd–95.8	[78]
	Portuguese seawaters	Nd–178	[80]
	Seawater from Singapore	<0.9–7	[90]
	Marina Bay, Singapore	13–30	[91]
	Durban Coast, South Africa	Nd–160	[92]
	Southern California Bight, USA	Nd–26	[85]
Fenoprofen	Gulf of Cadiz, Southwestern Spain	Nd–7.5	[78]
Ibuprofen	Seawater from Singapore	<2–9	[90]
	Marina Bay, Singapore	41–121	[91]
	Red Sea, Saudi Arabian coastal waters	508	[86]
	Santos Bay, Brazil	326–2094	[87]
	Durban Coast, South Africa	Nd–166	[92]
	San Francisco Bay, USA	12	[79]
	Southern California Bight, USA	Nd–30	[85]
	Portuguese seawaters	Nd–222	[80]
	Gulf of Cadiz, Southwestern Spain	Nd–1219.7	[78]
	Aegean and Dardanelles, Greece and Turkey	35	[79]
	Baltic Sea, Germany	109	[79]
	Seawater from Tromsø, Norway	Nd–0.7	[93]
	Mediterranean Sea, Israel	7.1	[79]
	Southwestern Taiwan	Nd–12.1	[81]
	Seawater from Northern Taiwan	<2.5–57.1	[89]

Nd: not detected.

Table 1. Concentrations of non-steroidal anti-inflammatory drugs in marine environments.

Non-Steroidal Anti-Inflammatory Drug	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Mefenamic acid	Mahdia Coast, Tunisia	Nd–0.6	[76]
	Eastern Mediterranean Sea, Greece	<0.2–11	[77]
	Gulf of Cadiz, Southwestern Spain	Nd–4.5	[78]
Phenylbutazone	Mahdia Coast, Tunisia	Nd–2	[76]
Phenazone	Gulf of Cadiz, Southwestern Spain	Nd–309.8	[78]
	Baltic Sea, Germany	5.9	[79]
	Aegean and Dardanelles, Greece and Turkey	2	[79]
Indomethacin	Gulf of Cadiz, Southwestern Spain	Nd–4.5	[78]
Tramadol	Eastern Mediterranean Sea, Greece	<0.1–1	[77]
Nimesulide	Portuguese seawaters	Nd–7.3	[80]
Codeine	Southwestern Taiwan	Nd–63.6	[81]
	Mediterranean coastal lagoon, Spain	1.8	[82]
Oxycodone	Mediterranean coastal lagoon, Spain	6.8	[82]
Acetylsalicylic acid	Portuguese seawaters	Nd–534	[80]
Acetaminophen	Gran Canaria Island, Spain	Nd–297	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–41.5	[78]
	Baltic Sea, Germany	48	[79]
	Aegean and Dardanelles, Greece and Turkey	2893	[79]
	Eastern Mediterranean Sea, Greece	<41	[77]
	Mediterranean Sea, Israel	12	[79]
	Seawater from Victoria BC, Canada	44.7	[84]
	San Francisco Bay, USA	85	[79]
	Southern California Bight, USA	Nd–11	[85]
	Portuguese seawaters	51–584	[80]
	Southwestern Taiwan	2.6–16.7	[81]
	Korean seawater	Nd–48	[81]
	Red Sea, Saudi Arabian coastal waters	2363	[86]
	Santos Bay, Brazil	Nd–34.6	[87]
Ketoprofen	Gran Canaria Island, Spain	Nd–106	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–2.6	[78]
	Southern Baltic Sea, Polish coastal zone	Nd–72.7	[88]
	Mahdia Coast, Tunisia	Nd–76	[76]
	Portuguese seawaters	10–90	[80]
	Southwestern Taiwan	Nd–23.3	[81]
	Northern Taiwan seawater	<1.7–6.59	[89]
Diclofenac	Gran Canaria Island, Spain	Nd–344	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–319	[78]
	Baltic Sea, Germany	9.2	[79]
	Mahdia Coast, Tunisia	Nd–23	[76]
	Eastern Mediterranean Sea, Greece	<1.4–16	[77]
	Mediterranean Sea, Israel	6.1	[79]
	Aegean and Dardanelles, Greece and Turkey	9.7	[79]
	Southern California Bight, USA	Nd–0.6	[85]
	Portuguese seawaters	Nd–241	[80]
	Santos Bay, Brazil	Nd–19.4	[87]
	Red Sea, Saudi Arabian coastal waters	14,020	[86]
	Seawater from Singapore	<2–12	[90]
	Marina Bay, Singapore	4–38	[91]
	Seawater from Northern Taiwan	<2.5–53.6	[89]
Naproxen	Eastern Mediterranean Sea, Greece	<0.01–0.8	[77]
	Gulf of Cadiz, South wstern Spain	Nd–95.8	[78]
	Portuguese seawaters	Nd–178	[80]
	Seawater from Singapore	<0.9–7	[90]
	Marina Bay, Singapore	13–30	[91]
	Durban Coast, South Africa	Nd–160	[92]
	Southern California Bight, USA	Nd–26	[85]
Fenoprofen	Gulf of Cadiz, Southwestern Spain	Nd–7.5	[78]
Ibuprofen	Seawater from Singapore	<2–9	[90]
	Marina Bay, Singapore	41–121	[91]
	Red Sea, Saudi Arabian coastal waters	508	[86]
	Santos Bay, Brazil	326–2094	[87]
	Durban Coast, South Africa	Nd–166	[92]
	San Francisco Bay, USA	12	[79]
	Southern California Bight, USA	Nd–30	[85]
	Portuguese seawaters	Nd–222	[80]
	Gulf of Cadiz, Southwestern Spain	Nd–1219.7	[78]
	Aegean and Dardanelles, Greece and Turkey	35	[79]
	Baltic Sea, Germany	109	[79]
	Seawater from Tromsø, Norway	Nd–0.7	[93]
	Mediterranean Sea, Israel	7.1	[79]
	Southwestern Taiwan	Nd–12.1	[81]
	Seawater from Northern Taiwan	<2.5–57.1	[89]

Nd: not detected.

Table 2. Concentrations of antibiotics in marine environments.

Antibiotic	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Amoxicillin	Eastern Mediterranean Sea, Greece	<128	[77]
Ampicillin	Southwest Taiwan	Nd–88.7	[81]
	Gulf of Cadiz, Southwestern Spain	Nd–2.0	[78]
Novobiocin	Gulf of Cadiz, Southwestern Spain	Nd–0.8	[78]
Clarithromycin	Eastern Mediterranean Sea, Greece	<1–1.5	[77]
	Aegean Sea and Dardanelles, Greece and Turkey	16	[79]
	Baltic Sea, Germany	14	[79]
	Gulf of Cadiz, Southwestern Spain	0.2–9.4	[78]
	Mediterranean coastal lagoon, Spain	9.6	[82]
	Pacific Ocean, USA	130	[79]
	Laizhou Bay, China	Nd–0.82	[94]
	Bohai Sea and Yellow Sea, China	Nd–0.51	[95]
	Yellow Sea, North China	2.6	[96]
Trimethoprim	Eastern Mediterranean Sea, Greece	<0.4–3	[77]
	Gulf of Cadiz, Southwestern Spain	Nd–10.6	[78]
	Mediterranean coastal lagoon, Spain	1.5	[82]
	Southern Baltic Sea, Polish coastal zone	Nd–2.9	[88]
	Mediterranean Sea, Tunisia	Nd–3500	[97]
	Southern Californi Bight, USA	Nd–2.1	[85]
Norfloxacin	Gran Canaria Island, Spain	Nd–3551	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–207.5	[78]
	Mediterranean Sea, Tunisia	Nd–20,700	[97]
	Yellow Sea Coast, China	Nd–109	[98]
	Laizhou Bay, China	7.5–103	[94]
	Bohai Sea, China	Nd–6800	[99]
	Victoria Harbour, Hong Kong	20.1	[100]
	Hong Kong coastal waters	Nd–8	[98]
	Korean seawater	Nd–0.512	[101]
Ciprofloxacin	Gran Canaria Island, Spain	Nd–303	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–211.7	[78]
	Yellow Sea Coast, China	Nd–26	[102]
	Laizhou Bay, China	Nd–66	[94]
	Bohai Bay, China	Nd–390	[99]
	Korean seawater	Nd–1.25	[101]
	Antarctica	4–128	[103]
Clindamycin	Antarctica	<0.1	[103]
	Gulf of Cadiz, Southwestern Spain	Nd–4.2	[78]
Enoxacin	Laizhou Bay, China	Nd–209	[94]
Ofloxacin	Bohai Bay, China	Nd–5100	[99]
	Laizhou Bay, China	Nd–6.5	[94]
	Victoria Harbour, Hong Kong	16.4	[24]
	Korean seawater	Nd–12.4	[101]
	Gulf of Cadiz, Southwestern Spain	Nd–34.4	[78]
Erythromycin	Mediterranean Sea, Southeast Spain	0.01–0.03	[104]
	Gulf of Cadiz, Southwestern Spain	Nd–2.3	[78]
	Mediterranean coastal lagoon, Spain	78.4	[82]
	Northern Adriatic Sea, Italy	5.8	[79]
	San Francisco Bay, USA	217	[79]
	Pacific Ocean, USA	86	[79]
	Mediterranean Sea, Tunisia	Nd–3900	[97]
	Bohai Bay, China	Nd–150	[99]
	Bohai Sea and Yellow Sea, China	0.13–6.7	[95]
	Yellow Sea, North China	25.2	[96]
	Laizhou Bay, China	0.9–8.5	[94]
	Hong Kong coastal waters	16–486	[98]
	Victoria Harbour, Hong Kong	5.2	[100]
	Korean seawater	Nd–0.196	[101]
	Southwestern Taiwan	Nd–26.6	[81]
Spiramycin	Gulf of Cadiz, Southwestern Spain	Nd–2.1	[78]
	Mediterranean Sea, Tunisia	Nd–66,400	[97]
	Korean seawater	Nd–7.24	[101]
Neospiramycin	Mediterranean Sea, Tunisia	Nd–4100	[97]
Josamycin	Mediterranean Sea, Tunisia	Nd–1500	[97]
Roxithromycin	Laizhou Bay, China	Nd–1.5	[94]
	Bohai Sea and Yellow Sea, China	Nd–0.26	[95]
	Yellow Sea, North China	6.9	[96]
	Victoria Harbour, Hong Kong	30.6	[105]
	Gulf of Cadiz, Southwestern Spain	Nd–1.3	[78]
	Baltic Sea, Germany	16	[79]
	Pacific Ocean, USA	141	[79]
Azithromycin	Laizhou Bay, China	Nd–1.2	[94]
	Bohai Sea and Yellow Sea, China	Nd–0.39	[95]
	Yellow Sea, North China	2.5	[96]
	Gulf of Cadiz, Southwestern Spain	Nd–17.8	[78]
	Mediterranean coastal lagoon, Spain	163.8	[82]
Lomefloxacin	Yellow Sea Coast, China	Nd–1.2	[102]
Danofloxacin	Yellow Sea Coast, China	Nd–30	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–157.5	[78]
Enrofloxacin	Yellow Sea Coast, China	0.78–5.1	[102]
	Laizhou Bay, China	Nd–7.6	[94]
	Gulf of Cadiz, Southwestern Spain	Nd–122	[78]
	Southern Baltic Sea, Polish coastal zone	Nd	[88]
	Mediterranean Sea, Tunisia	Nd–40,200	[97]
Marbofloxacin	Yellow Sea Coast, China	Nd–22	[102]
Fleorxacin	Yellow Sea Coast, China	Nd–1.4	[102]
Orbifloxacin	Yellow Sea Coast, China	Nd–2.7	[102]
Difloxacin	Yellow Sea Coast, China	Nd–20.7	[102]
Sarafloxacin	Yellow Sea Coast, China	Nd–14.6	[102]
	Mediterranean Sea, Tunisia	Nd–5300	[97]
Sparfloxacin	Yellow Sea Coast, China	Nd–0.79	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–14.9	[78]
Lincomycin	Korean seawater	Nd–438	[101]
	Gulf of Cadiz, Southwestern Spain	Nd–6.1	[78]
Cefalexin	Southwestern Taiwan	Nd–9.19	[81]
	Hong Kong coastal waters	Nd–182	[98]
Cefaclor	Gulf of Cadiz, Southwestern Spain	Nd–9.4	[78]
Cefdinir	Gulf of Cadiz, Southwestern Spain	Nd–15.8	[78]
Cefquinone	Gulf of Cadiz, Southwestern Spain	Nd–44.9	[78]
Ceftiotur	Gulf of Cadiz, Southwestern Spain	Nd–1.7	[78]
Sulfadiazine	Eastern Mediterranean Sea, Greece	<0.1–2	[77]
	Gulf of Cadiz, Southwestern Spain	Nd–1.8	[78]
	Mediterranean Sea, Tunisia	Nd–29,100	[97]
	Mahdia Coasta, Tunisia	6–11	[76]
	Dalian Coast, China	Nd–2	[106]
	Yellow Sea, North China	0.24	[96]
	Yellow Sea Coast, China	Nd–3.0	[102]
	Bohai Bay, China	Nd–41	[99]
	Bohai Sea and Yellow Sea, China	Nd–0.36	[95]
	Laizhou Bay, China	Nd–0.43	[94]
Sulfamerazine	Mediterranean Sea, Tunisia	Nd–4500	[97]
	Southern Baltic Sea, Polish coastal zone	Nd	[88]
Sulfamoxole	Mediterranean Sea, Tunisia	Nd–800	[97]
Sulfamethoxazole	Eastern Mediterranean Sea, Greece	<0.1–6	[77]
	Aegean Sea and Dardanelles, Greece and Turkey	11	[79]
	Northern Adriatic Sea, Italy	4.1	[79]
	Mediterranean Sea, Tunisia	Nd–2400	[97]
	Gulf of Cadiz, Southwestern Spain	Nd–99	[78]
	Mediterranean coastal lagoon, Spain	94	[82]
	Baltic Sea, Germany	42	[79]
	Southern Baltic Sea, Polish coastal zone	Nd–20.0	[88]
	Baltic Sea, Poland	Nd–10.8	[88]
	Mahdia Coast, Tunisia	2–6	[76]
	Red Sea, Saudi Arabian coastal waters	63	[86]
	Dalian Coast, China	Nd–2.2	[106]
	Yellow Sea Coast, China	Nd–212	[102]
	Bohai Bay, China	Nd–140	[99]
	Bohai Sea and Yellow Sea, China	Nd	[95]
	Yellow Sea, North China	50.4	[96]
	Laizhou Bay, China	1.5–82	[94]
	Korean seawater	Nd–2.20	[101]
	San Francisco Bay, USA	61	[79]
	Pacific Ocean, USA	6.4	[79]
	Southern California Bight, USA	Nd–3.4	[85]
	German Baltic Sea	1.5	[107]
Sulfathiazole	Mahdia Coast, Tunisia	Nd–3	[76]
	Bohai Sea and Yellow Sea, China	Nd–0.17	[95]
	Dalian Coast, China	Nd–1.2	[106]
	Korean seawater	7.01–18.6	[101]
Sulfaphenazole	Mediterranean Sea, Tunisia	Nd–600	[97]
Sulfamethizole	Mahdia Coast, Tunisia	4–11	[76]
	Mediterranean Sea, Tunisia	Nd–2800	[97]
	Dalian Coast, China	Nd–13	[106]
	Gulf of Cadiz, Southwestern Spain	Nd–67.1	[78]
Metronidazole	Mediterranean Sea, Southeast Spain	13.4	[104]
	Gulf of Cadiz, Southwestern Spain	Nd–2.3	[78]
Nitrofurantoin	Gulf of Cadiz, Southwestern Spain	Nd–21.7	[78]
Ornidazole	Gulf of Cadiz, Southwestern Spain	Nd–1.9	[78]
Sulfamethazine	Mahdia Coast, Tunisia	Nd–3	[76]
	Dalian Coast, China	Nd–2.8	[106]
	Yellow Sea Coast, China	Nd–37	[102]
	Bohai Bay, China	Nd–130	[99]
	Laizhou Bay, China	Nd–1.5	[94]
	Gulf of Cadiz, Southwestern Spain	Nd–9.1	[78]
	Southern Baltic Sea, Polish coastal zone	Nd	[88]
Sulfadimidine	Bohai Sea and Yellow Sea, China	Nd–0.16	[95]
	Yellow Sea, North China	0.35	[96]
	Mediterranean Sea, Tunisia	Nd–1800	[97]
Sulfaquinoxaline	Mediterranean Sea, Tunisia	Nd–1900	[97]
Sulfaguanidine	Mediterranean Sea, Tunisia	Nd–200	[97]
Sulfamethoxyrpyridazine	Mahdia Coast, Tunisia	Nd–5	[76]
Sulfacetamide	Dalian Coast, China	Nd–1.5	[106]
	Yellow Sea Coast, China	Nd–4.3	[102]
	Bohai Sea and Yellow Sea, China	Nd–0.12	[95]
Sulfameter	Dalian Coast, China	Nd–1.9	[106]
	Yellow Sea Coast, China	Nd–1.2	[102]
Sulfamonomethoxine	Dalian Coast, China	Nd–2.3	[106]
	Yellow Sea Coast, China	Nd–4.6	[102]
Sulfadimethoxine	Dalian Coast, China	Nd–1.9	[106]
	Yellow Sea Coast, China	Nd–1.9	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–0.9	[78]
	Southern Baltic Sea, Polish coastal zone	Nd–1.0	[88]
	Baltic Sea, Poland	Nd–0.8	[88]
Sulfapyridine	Mediterranean Sea, Tunisia	Nd–400	[97]
	Southern Baltic Sea, Polish coastal zone	Nd–33.2	[88]
Chloramphenicol	Dalian Coast, China	Nd–1.4	[106]
	Gulf of Cadiz, Southwestern Spain	Nd–8.1	[78]
Timulin	Gulf of Cadiz, Southwestern Spain	Nd–0.8	[78]
Florophenicol	Dalian Coast, China	Nd–2.3	[106]
Oxytetracycline	Dalian Coast, China	1.1–6.3	[106]
	Yellow Sea Coast, China	Nd–13.0	[102]
	Bohai Bay, China	Nd–270	[99]
	Gulf of Cadiz, Southwestern Spain	Nd–25.1	[78]
Doxycycline	Dalian Coast, China	Nd–1.6	[106]
	Yellow Sea Coast, China	Nd–3.2	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–10.3	[78]
Tetracycline	Dalian Coast, China	Nd–3.8	[106]
	Yellow Sea Coast, China	Nd–5.3	[102]
	Bohai Bay, China	Nd–30	[99]
	Hong Kong coastal waters	Nd–122	[98]
	Gulf of Cadiz, Southwestern Spain	Nd–63.3	[78]
Sulfisoxazole	Yellow Sea Coast, China	Nd–16.5	[102]
	Mediterranean Sea, Tunisia	100–700	[97]
Sulfachloropyridazine	Yellow Sea Coast, China	Nd–5.9	[102]
Oxolinic acid	Yellow Sea Coast, China	29–105	[102]
Pyrrole acid	Yellow Sea Coast, China	0.95–17.5	[102]
Nalidixic acid	Yellow Sea Coast, China	Nd–28.9	[102]
	Mediterranean Sea, Tunisia	Nd–16,700	[97]
Pefloxacic	Yellow Sea Coast, China	Nd–14.6	[102]
Flumequine	Yellow Sea Coast, China	Nd–7.0	[102]
	Mediterranean Sea, Southeast Spain	0.13	[104]
	Gulf of Cadiz, Southwestern Spain	Nd–3.6	[78]
Dapsone	Mediterranean Sea, Tunisia	Nd–2800	[97]

Nd: not detected.

Table 2. Concentrations of antibiotics in marine environments.

Antibiotic	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Amoxicillin	Eastern Mediterranean Sea, Greece	<128	[77]
Ampicillin	Southwest Taiwan	Nd–88.7	[81]
	Gulf of Cadiz, Southwestern Spain	Nd–2.0	[78]
Novobiocin	Gulf of Cadiz, Southwestern Spain	Nd–0.8	[78]
Clarithromycin	Eastern Mediterranean Sea, Greece	<1–1.5	[77]
	Aegean Sea and Dardanelles, Greece and Turkey	16	[79]
	Baltic Sea, Germany	14	[79]
	Gulf of Cadiz, Southwestern Spain	0.2–9.4	[78]
	Mediterranean coastal lagoon, Spain	9.6	[82]
	Pacific Ocean, USA	130	[79]
	Laizhou Bay, China	Nd–0.82	[94]
	Bohai Sea and Yellow Sea, China	Nd–0.51	[95]
	Yellow Sea, North China	2.6	[96]
Trimethoprim	Eastern Mediterranean Sea, Greece	<0.4–3	[77]
	Gulf of Cadiz, Southwestern Spain	Nd–10.6	[78]
	Mediterranean coastal lagoon, Spain	1.5	[82]
	Southern Baltic Sea, Polish coastal zone	Nd–2.9	[88]
	Mediterranean Sea, Tunisia	Nd–3500	[97]
	Southern Californi Bight, USA	Nd–2.1	[85]
Norfloxacin	Gran Canaria Island, Spain	Nd–3551	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–207.5	[78]
	Mediterranean Sea, Tunisia	Nd–20,700	[97]
	Yellow Sea Coast, China	Nd–109	[98]
	Laizhou Bay, China	7.5–103	[94]
	Bohai Sea, China	Nd–6800	[99]
	Victoria Harbour, Hong Kong	20.1	[100]
	Hong Kong coastal waters	Nd–8	[98]
	Korean seawater	Nd–0.512	[101]
Ciprofloxacin	Gran Canaria Island, Spain	Nd–303	[83]
	Gulf of Cadiz, Southwestern Spain	Nd–211.7	[78]
	Yellow Sea Coast, China	Nd–26	[102]
	Laizhou Bay, China	Nd–66	[94]
	Bohai Bay, China	Nd–390	[99]
	Korean seawater	Nd–1.25	[101]
	Antarctica	4–128	[103]
Clindamycin	Antarctica	<0.1	[103]
	Gulf of Cadiz, Southwestern Spain	Nd–4.2	[78]
Enoxacin	Laizhou Bay, China	Nd–209	[94]
Ofloxacin	Bohai Bay, China	Nd–5100	[99]
	Laizhou Bay, China	Nd–6.5	[94]
	Victoria Harbour, Hong Kong	16.4	[24]
	Korean seawater	Nd–12.4	[101]
	Gulf of Cadiz, Southwestern Spain	Nd–34.4	[78]
Erythromycin	Mediterranean Sea, Southeast Spain	0.01–0.03	[104]
	Gulf of Cadiz, Southwestern Spain	Nd–2.3	[78]
	Mediterranean coastal lagoon, Spain	78.4	[82]
	Northern Adriatic Sea, Italy	5.8	[79]
	San Francisco Bay, USA	217	[79]
	Pacific Ocean, USA	86	[79]
	Mediterranean Sea, Tunisia	Nd–3900	[97]
	Bohai Bay, China	Nd–150	[99]
	Bohai Sea and Yellow Sea, China	0.13–6.7	[95]
	Yellow Sea, North China	25.2	[96]
	Laizhou Bay, China	0.9–8.5	[94]
	Hong Kong coastal waters	16–486	[98]
	Victoria Harbour, Hong Kong	5.2	[100]
	Korean seawater	Nd–0.196	[101]
	Southwestern Taiwan	Nd–26.6	[81]
Spiramycin	Gulf of Cadiz, Southwestern Spain	Nd–2.1	[78]
	Mediterranean Sea, Tunisia	Nd–66,400	[97]
	Korean seawater	Nd–7.24	[101]
Neospiramycin	Mediterranean Sea, Tunisia	Nd–4100	[97]
Josamycin	Mediterranean Sea, Tunisia	Nd–1500	[97]
Roxithromycin	Laizhou Bay, China	Nd–1.5	[94]
	Bohai Sea and Yellow Sea, China	Nd–0.26	[95]
	Yellow Sea, North China	6.9	[96]
	Victoria Harbour, Hong Kong	30.6	[105]
	Gulf of Cadiz, Southwestern Spain	Nd–1.3	[78]
	Baltic Sea, Germany	16	[79]
	Pacific Ocean, USA	141	[79]
Azithromycin	Laizhou Bay, China	Nd–1.2	[94]
	Bohai Sea and Yellow Sea, China	Nd–0.39	[95]
	Yellow Sea, North China	2.5	[96]
	Gulf of Cadiz, Southwestern Spain	Nd–17.8	[78]
	Mediterranean coastal lagoon, Spain	163.8	[82]
Lomefloxacin	Yellow Sea Coast, China	Nd–1.2	[102]
Danofloxacin	Yellow Sea Coast, China	Nd–30	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–157.5	[78]
Enrofloxacin	Yellow Sea Coast, China	0.78–5.1	[102]
	Laizhou Bay, China	Nd–7.6	[94]
	Gulf of Cadiz, Southwestern Spain	Nd–122	[78]
	Southern Baltic Sea, Polish coastal zone	Nd	[88]
	Mediterranean Sea, Tunisia	Nd–40,200	[97]
Marbofloxacin	Yellow Sea Coast, China	Nd–22	[102]
Fleorxacin	Yellow Sea Coast, China	Nd–1.4	[102]
Orbifloxacin	Yellow Sea Coast, China	Nd–2.7	[102]
Difloxacin	Yellow Sea Coast, China	Nd–20.7	[102]
Sarafloxacin	Yellow Sea Coast, China	Nd–14.6	[102]
	Mediterranean Sea, Tunisia	Nd–5300	[97]
Sparfloxacin	Yellow Sea Coast, China	Nd–0.79	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–14.9	[78]
Lincomycin	Korean seawater	Nd–438	[101]
	Gulf of Cadiz, Southwestern Spain	Nd–6.1	[78]
Cefalexin	Southwestern Taiwan	Nd–9.19	[81]
	Hong Kong coastal waters	Nd–182	[98]
Cefaclor	Gulf of Cadiz, Southwestern Spain	Nd–9.4	[78]
Cefdinir	Gulf of Cadiz, Southwestern Spain	Nd–15.8	[78]
Cefquinone	Gulf of Cadiz, Southwestern Spain	Nd–44.9	[78]
Ceftiotur	Gulf of Cadiz, Southwestern Spain	Nd–1.7	[78]
Sulfadiazine	Eastern Mediterranean Sea, Greece	<0.1–2	[77]
	Gulf of Cadiz, Southwestern Spain	Nd–1.8	[78]
	Mediterranean Sea, Tunisia	Nd–29,100	[97]
	Mahdia Coasta, Tunisia	6–11	[76]
	Dalian Coast, China	Nd–2	[106]
	Yellow Sea, North China	0.24	[96]
	Yellow Sea Coast, China	Nd–3.0	[102]
	Bohai Bay, China	Nd–41	[99]
	Bohai Sea and Yellow Sea, China	Nd–0.36	[95]
	Laizhou Bay, China	Nd–0.43	[94]
Sulfamerazine	Mediterranean Sea, Tunisia	Nd–4500	[97]
	Southern Baltic Sea, Polish coastal zone	Nd	[88]
Sulfamoxole	Mediterranean Sea, Tunisia	Nd–800	[97]
Sulfamethoxazole	Eastern Mediterranean Sea, Greece	<0.1–6	[77]
	Aegean Sea and Dardanelles, Greece and Turkey	11	[79]
	Northern Adriatic Sea, Italy	4.1	[79]
	Mediterranean Sea, Tunisia	Nd–2400	[97]
	Gulf of Cadiz, Southwestern Spain	Nd–99	[78]
	Mediterranean coastal lagoon, Spain	94	[82]
	Baltic Sea, Germany	42	[79]
	Southern Baltic Sea, Polish coastal zone	Nd–20.0	[88]
	Baltic Sea, Poland	Nd–10.8	[88]
	Mahdia Coast, Tunisia	2–6	[76]
	Red Sea, Saudi Arabian coastal waters	63	[86]
	Dalian Coast, China	Nd–2.2	[106]
	Yellow Sea Coast, China	Nd–212	[102]
	Bohai Bay, China	Nd–140	[99]
	Bohai Sea and Yellow Sea, China	Nd	[95]
	Yellow Sea, North China	50.4	[96]
	Laizhou Bay, China	1.5–82	[94]
	Korean seawater	Nd–2.20	[101]
	San Francisco Bay, USA	61	[79]
	Pacific Ocean, USA	6.4	[79]
	Southern California Bight, USA	Nd–3.4	[85]
	German Baltic Sea	1.5	[107]
Sulfathiazole	Mahdia Coast, Tunisia	Nd–3	[76]
	Bohai Sea and Yellow Sea, China	Nd–0.17	[95]
	Dalian Coast, China	Nd–1.2	[106]
	Korean seawater	7.01–18.6	[101]
Sulfaphenazole	Mediterranean Sea, Tunisia	Nd–600	[97]
Sulfamethizole	Mahdia Coast, Tunisia	4–11	[76]
	Mediterranean Sea, Tunisia	Nd–2800	[97]
	Dalian Coast, China	Nd–13	[106]
	Gulf of Cadiz, Southwestern Spain	Nd–67.1	[78]
Metronidazole	Mediterranean Sea, Southeast Spain	13.4	[104]
	Gulf of Cadiz, Southwestern Spain	Nd–2.3	[78]
Nitrofurantoin	Gulf of Cadiz, Southwestern Spain	Nd–21.7	[78]
Ornidazole	Gulf of Cadiz, Southwestern Spain	Nd–1.9	[78]
Sulfamethazine	Mahdia Coast, Tunisia	Nd–3	[76]
	Dalian Coast, China	Nd–2.8	[106]
	Yellow Sea Coast, China	Nd–37	[102]
	Bohai Bay, China	Nd–130	[99]
	Laizhou Bay, China	Nd–1.5	[94]
	Gulf of Cadiz, Southwestern Spain	Nd–9.1	[78]
	Southern Baltic Sea, Polish coastal zone	Nd	[88]
Sulfadimidine	Bohai Sea and Yellow Sea, China	Nd–0.16	[95]
	Yellow Sea, North China	0.35	[96]
	Mediterranean Sea, Tunisia	Nd–1800	[97]
Sulfaquinoxaline	Mediterranean Sea, Tunisia	Nd–1900	[97]
Sulfaguanidine	Mediterranean Sea, Tunisia	Nd–200	[97]
Sulfamethoxyrpyridazine	Mahdia Coast, Tunisia	Nd–5	[76]
Sulfacetamide	Dalian Coast, China	Nd–1.5	[106]
	Yellow Sea Coast, China	Nd–4.3	[102]
	Bohai Sea and Yellow Sea, China	Nd–0.12	[95]
Sulfameter	Dalian Coast, China	Nd–1.9	[106]
	Yellow Sea Coast, China	Nd–1.2	[102]
Sulfamonomethoxine	Dalian Coast, China	Nd–2.3	[106]
	Yellow Sea Coast, China	Nd–4.6	[102]
Sulfadimethoxine	Dalian Coast, China	Nd–1.9	[106]
	Yellow Sea Coast, China	Nd–1.9	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–0.9	[78]
	Southern Baltic Sea, Polish coastal zone	Nd–1.0	[88]
	Baltic Sea, Poland	Nd–0.8	[88]
Sulfapyridine	Mediterranean Sea, Tunisia	Nd–400	[97]
	Southern Baltic Sea, Polish coastal zone	Nd–33.2	[88]
Chloramphenicol	Dalian Coast, China	Nd–1.4	[106]
	Gulf of Cadiz, Southwestern Spain	Nd–8.1	[78]
Timulin	Gulf of Cadiz, Southwestern Spain	Nd–0.8	[78]
Florophenicol	Dalian Coast, China	Nd–2.3	[106]
Oxytetracycline	Dalian Coast, China	1.1–6.3	[106]
	Yellow Sea Coast, China	Nd–13.0	[102]
	Bohai Bay, China	Nd–270	[99]
	Gulf of Cadiz, Southwestern Spain	Nd–25.1	[78]
Doxycycline	Dalian Coast, China	Nd–1.6	[106]
	Yellow Sea Coast, China	Nd–3.2	[102]
	Gulf of Cadiz, Southwestern Spain	Nd–10.3	[78]
Tetracycline	Dalian Coast, China	Nd–3.8	[106]
	Yellow Sea Coast, China	Nd–5.3	[102]
	Bohai Bay, China	Nd–30	[99]
	Hong Kong coastal waters	Nd–122	[98]
	Gulf of Cadiz, Southwestern Spain	Nd–63.3	[78]
Sulfisoxazole	Yellow Sea Coast, China	Nd–16.5	[102]
	Mediterranean Sea, Tunisia	100–700	[97]
Sulfachloropyridazine	Yellow Sea Coast, China	Nd–5.9	[102]
Oxolinic acid	Yellow Sea Coast, China	29–105	[102]
Pyrrole acid	Yellow Sea Coast, China	0.95–17.5	[102]
Nalidixic acid	Yellow Sea Coast, China	Nd–28.9	[102]
	Mediterranean Sea, Tunisia	Nd–16,700	[97]
Pefloxacic	Yellow Sea Coast, China	Nd–14.6	[102]
Flumequine	Yellow Sea Coast, China	Nd–7.0	[102]
	Mediterranean Sea, Southeast Spain	0.13	[104]
	Gulf of Cadiz, Southwestern Spain	Nd–3.6	[78]
Dapsone	Mediterranean Sea, Tunisia	Nd–2800	[97]

Nd: not detected.

Table 3. Concentrations of antidepressant drugs in marine environments.

Antidepressant Drug	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Carbamazepine	Eastern Mediterranean Sea, Greece	<1.4	[77]
	French coast on the Mediterranean Sea	0.05–0.71	[108]
	Mediterranean coastal lagoon, Spain	4.9	[82]
	Mediterranean Sea, Israel	8.8	[79]
	Gulf of Cadiz, Southwestern Spain	Nd–31.1	[78]
	North Seawater, Germany	2	[109]
	Baltic Sea, Germany	157	[79]
	Aegean and Dardanelles, Greece and Turkey	22	[79]
	Northern Adriatic Sea, Italy	3.1	[79]
	Southwestern Taiwan	Nd–3.83	[81]
	Red Sea, Saudi Arabian coastal waters	110	[86]
	San Francisco Bay, USA	13	[79]
	Southern California Bight, USA	Nd–0.9	[85]
	Californian Coast, USA	Nd–21	[110]
	Mahdia Coast, Tunisia	Nd–0.5	[76]
	Korean seawater	4.58–38.6	[101]
	Seawater from Singapore	<0.3–11	[90]
Norvenlafaxine	Eastern Mediterranean Sea, Greece	<0.01–2	[77]
Venlafaxine	Rías Baixas coastline, Northwestern Spain	Nd–291	[111]
Citalopram	Eastern Mediterranean Sea, Greece	<0.06–8	[77]
	Rías Baixas coastline, Northwestern Spain	Nd–92.5	[111]
	Mediterranean Sea, Israel	4.3	[79]
	Pacific Ocean, USA	27	[79]
Fluoxetine	Mahdia Coast, Tunisia	Nd–41	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–0.6	[78]
	Rías Baixas coastline, Northwestern Spain	Nd–10.6	[111]
	Pacific Ocean, USA	90	[79]
Amitriptyline	Mahdia Coast, Tunisia	Nd–10	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–0.4	[78]
Hydroxyzine	Rías Baixas coastline, Northwestern Spain	Nd–0.57	[111]
Sertraline	Rías Baixas coastline, Northwestern Spain	Nd–15.3	[111]

Nd: not detected.

Table 3. Concentrations of antidepressant drugs in marine environments.

Antidepressant Drug	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Carbamazepine	Eastern Mediterranean Sea, Greece	<1.4	[77]
	French coast on the Mediterranean Sea	0.05–0.71	[108]
	Mediterranean coastal lagoon, Spain	4.9	[82]
	Mediterranean Sea, Israel	8.8	[79]
	Gulf of Cadiz, Southwestern Spain	Nd–31.1	[78]
	North Seawater, Germany	2	[109]
	Baltic Sea, Germany	157	[79]
	Aegean and Dardanelles, Greece and Turkey	22	[79]
	Northern Adriatic Sea, Italy	3.1	[79]
	Southwestern Taiwan	Nd–3.83	[81]
	Red Sea, Saudi Arabian coastal waters	110	[86]
	San Francisco Bay, USA	13	[79]
	Southern California Bight, USA	Nd–0.9	[85]
	Californian Coast, USA	Nd–21	[110]
	Mahdia Coast, Tunisia	Nd–0.5	[76]
	Korean seawater	4.58–38.6	[101]
	Seawater from Singapore	<0.3–11	[90]
Norvenlafaxine	Eastern Mediterranean Sea, Greece	<0.01–2	[77]
Venlafaxine	Rías Baixas coastline, Northwestern Spain	Nd–291	[111]
Citalopram	Eastern Mediterranean Sea, Greece	<0.06–8	[77]
	Rías Baixas coastline, Northwestern Spain	Nd–92.5	[111]
	Mediterranean Sea, Israel	4.3	[79]
	Pacific Ocean, USA	27	[79]
Fluoxetine	Mahdia Coast, Tunisia	Nd–41	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–0.6	[78]
	Rías Baixas coastline, Northwestern Spain	Nd–10.6	[111]
	Pacific Ocean, USA	90	[79]
Amitriptyline	Mahdia Coast, Tunisia	Nd–10	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–0.4	[78]
Hydroxyzine	Rías Baixas coastline, Northwestern Spain	Nd–0.57	[111]
Sertraline	Rías Baixas coastline, Northwestern Spain	Nd–15.3	[111]

Nd: not detected.

Table 4. Concentration of cardiovascular drugs in marine environments.

Cardiovascular Drug	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Timolol	Mahdia Coast, Tunisia	Nd–0.3	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–1.1	[78]
Nadolol	Mahdia Coast, Tunisia	Nd–0.8	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–1.6	[78]
Atenolol	Southern California Bight, USA	Nd–11	[85]
	Korean seawater	Nd–85.7	[101]
	Gulf of Cadiz, Southwestern Spain	0.4–138.9	[78]
	Baltic Sea, Germany	13	[79]
	Aegean and Dardanelles, Greece and Turkey	194	[79]
	San Francisco Bay, USA	57	[79]
	Santos Bay, Brazil	Nd	[87]
Propanolol	Korean seawater	Nd–11.9	[101]
	Mediterranean coastal lagoon, Spain	0.5	[82]
	Gulf of Cadiz, Southwestern Spain	Nd–5.9	[78]
Metoprolol	Mediterranean coastal lagoon, Spain	0.73	[82]
	Gulf of Cadiz, Southwestern Spain	Nd–5.1	[78]
	Baltic Sea, Germany	158	[79]
	Aegean and Dardanelles, Greece and Turkey	6	[79]
	San Francisco Bay, USA	32	[79]
	Mediterranean Sea, Israel	6.7	[79]
Sotalol	Mediterranean coastal lagoon, Spain	0.8	[82]
	Baltic Sea, Germany	65	[79]
	Aegean and Dardanelles, Greece and Turkey	67	[79]
	San Francisco Bay, USA	12	[79]
Pindolol	Gulf of Cadiz, Southwestern Spain	Nd–0.7	[78]

Nd: not detected.

Table 4. Concentration of cardiovascular drugs in marine environments.

Cardiovascular Drug	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Timolol	Mahdia Coast, Tunisia	Nd–0.3	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–1.1	[78]
Nadolol	Mahdia Coast, Tunisia	Nd–0.8	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–1.6	[78]
Atenolol	Southern California Bight, USA	Nd–11	[85]
	Korean seawater	Nd–85.7	[101]
	Gulf of Cadiz, Southwestern Spain	0.4–138.9	[78]
	Baltic Sea, Germany	13	[79]
	Aegean and Dardanelles, Greece and Turkey	194	[79]
	San Francisco Bay, USA	57	[79]
	Santos Bay, Brazil	Nd	[87]
Propanolol	Korean seawater	Nd–11.9	[101]
	Mediterranean coastal lagoon, Spain	0.5	[82]
	Gulf of Cadiz, Southwestern Spain	Nd–5.9	[78]
Metoprolol	Mediterranean coastal lagoon, Spain	0.73	[82]
	Gulf of Cadiz, Southwestern Spain	Nd–5.1	[78]
	Baltic Sea, Germany	158	[79]
	Aegean and Dardanelles, Greece and Turkey	6	[79]
	San Francisco Bay, USA	32	[79]
	Mediterranean Sea, Israel	6.7	[79]
Sotalol	Mediterranean coastal lagoon, Spain	0.8	[82]
	Baltic Sea, Germany	65	[79]
	Aegean and Dardanelles, Greece and Turkey	67	[79]
	San Francisco Bay, USA	12	[79]
Pindolol	Gulf of Cadiz, Southwestern Spain	Nd–0.7	[78]

Nd: not detected.

Table 5. Concentration of lipid regulators in marine environments.

Lipid Regulator	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Fenofibrate	Mahdia Coast, Tunisia	Nd–14	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–1.1	[78]
Bezafibrate	Gulf of Cadiz, Southwestern Spain	Nd–0.5	[78]
	Aegean and Dardanelles, Greece and Turkey	3.5	[79]
	Mediterranean Sea, Israel	3.8	[79]
Gemfibrozil	Seawater from Singapore	<0.09–20	[90]
	Marina Bay, Singapore	1–9	[91]
	Pacific Ocean, USA	6.2	[79]
	San Francisco Bay, USA	43	[79]
	Southern California Bight, USA	Nd–13	[85]
	Southwestern Taiwan	Nd–3.67	[81]
	Mediterranean coastal lagoon, Spain	3.3	[82]
	Gulf of Cadiz, Southwestern Spain	Nd–5.7	[78]
	Aegean and Dardanelles, Greece and Turkey	18	[79]
Atorvastatin	Southern California Bight, USA	Nd–0.4	[85]

Nd: not detected.

Table 5. Concentration of lipid regulators in marine environments.

Lipid Regulator	Environment Studied	Concentration in Water Sample (ng/L)	Reference
Fenofibrate	Mahdia Coast, Tunisia	Nd–14	[76]
	Gulf of Cadiz, Southwestern Spain	Nd–1.1	[78]
Bezafibrate	Gulf of Cadiz, Southwestern Spain	Nd–0.5	[78]
	Aegean and Dardanelles, Greece and Turkey	3.5	[79]
	Mediterranean Sea, Israel	3.8	[79]
Gemfibrozil	Seawater from Singapore	<0.09–20	[90]
	Marina Bay, Singapore	1–9	[91]
	Pacific Ocean, USA	6.2	[79]
	San Francisco Bay, USA	43	[79]
	Southern California Bight, USA	Nd–13	[85]
	Southwestern Taiwan	Nd–3.67	[81]
	Mediterranean coastal lagoon, Spain	3.3	[82]
	Gulf of Cadiz, Southwestern Spain	Nd–5.7	[78]
	Aegean and Dardanelles, Greece and Turkey	18	[79]
Atorvastatin	Southern California Bight, USA	Nd–0.4	[85]

Nd: not detected.

Table 6. Concentration of EE2 in marine environments.

Estrogen	Environment Studies	Concentration in Water Sample (ng/L)	Reference
17α ethynylestradiol (EE2)	Aegean Sea, Greece	Nd	[112]
	Singapore	Nd	[90]
	Baltic Sea	1.7–8.0	[113]
	Southeastern Australia	<0.20	[114]
	Suruga Bay, Japan	<3	[115]
	Sein River, France	Nd	[116]
	Bahia, Brazil	Nd	[117]
	Yangtze Estuary, China	Nd	[118]
	Scheldt Estuary, Belgium	Nd	[119]
	Venice Lagoon, Italy	<5–75	[120]
	Venice Lagoon, Italy	<0.8–34	[121]
	Douro River Estuary, Portugal	Nd–83.1	[122]
	Mondego River Estuary, Portugal	Nd	[123]
	Halifax Harbour, Canada	Nd–0.14	[124]
	Mondego River, Portugal	4	[125]
	Ria de Aveiro, Portugal	20.7–33.2	[126]
	Sado River Estuary, Portugal	1.1–3.3	[125]
	Dublin Bay, Ireland	Nd	[127]
	Yangtze River Estuary, China	Nd–0.11	[128]
	Rio de la Plata Estuary, Argentina	Nd–43	[129]
	Hong Kong and Japan	Nd	[130]
	Yundang Lagoon, China	Nd–0.43	[131]

Nd: not detected.

Table 6. Concentration of EE2 in marine environments.

Estrogen	Environment Studies	Concentration in Water Sample (ng/L)	Reference
17α ethynylestradiol (EE2)	Aegean Sea, Greece	Nd	[112]
	Singapore	Nd	[90]
	Baltic Sea	1.7–8.0	[113]
	Southeastern Australia	<0.20	[114]
	Suruga Bay, Japan	<3	[115]
	Sein River, France	Nd	[116]
	Bahia, Brazil	Nd	[117]
	Yangtze Estuary, China	Nd	[118]
	Scheldt Estuary, Belgium	Nd	[119]
	Venice Lagoon, Italy	<5–75	[120]
	Venice Lagoon, Italy	<0.8–34	[121]
	Douro River Estuary, Portugal	Nd–83.1	[122]
	Mondego River Estuary, Portugal	Nd	[123]
	Halifax Harbour, Canada	Nd–0.14	[124]
	Mondego River, Portugal	4	[125]
	Ria de Aveiro, Portugal	20.7–33.2	[126]
	Sado River Estuary, Portugal	1.1–3.3	[125]
	Dublin Bay, Ireland	Nd	[127]
	Yangtze River Estuary, China	Nd–0.11	[128]
	Rio de la Plata Estuary, Argentina	Nd–43	[129]
	Hong Kong and Japan	Nd	[130]
	Yundang Lagoon, China	Nd–0.43	[131]

Nd: not detected.

2.2. Occurrence of Pharmaceuticals in the Marine Environment

The fate and persistence of pharmaceuticals in marine ecosystems are influenced by the physicochemical characteristics of the active principles, physicochemical processes, and environmental factors [1,132]. Once released into the marine environment, pharmaceuticals can be subjected to biotic and abiotic transformation, sorbed to suspended particulate matter, or, in some cases, accumulated in the tissues of marine organisms [133]. The principal processes involved in the transformation of pharmaceuticals in the marine environment are aerobic and anaerobic biodegradation, abiotic transformation (dilution and movement within the aquatic milieu) through degradation by UV light and hydrolysis [134]. Furthermore, the dissimilarities in the environmental conditions, including temperature, salinity, pH, and organic matter, can influence these processes and the fate and persistence of pharmaceuticals in seawater [109]. It has been demonstrated that pH and salinity can affect the electrostatic characteristics of pharmaceuticals, resulting in the generation of numerous ionizable functional groups with varying acid dissociation constants (pKa). In particular, changes in pH determine the degree of ionisation and can affect the lipophilicity of pharmaceuticals, rendering these compounds more effectively adsorbed onto particulate matter at pH 8 [29,135,136]. Similarly, salinity exerts an influence on the persistence of pharmaceuticals in marine ecosystems by acting as a natural filter and causing a reduction in the pharmaceuticals’ solubility in seawater. For example, the amount of gemfibrozil and mefenamic acid can be assessed from marine surface waters due to the high content of salt [90,137]. Furthermore, the intensity of solar radiation may influence the behavior of pharmaceuticals in marine environments, leading to the photolysis of certain contaminants such as the antiepileptic medication carbamazepine into acridine, a substance known for its high toxicity, mutagenic properties, and carcinogenic effects [138,139].

3. Biological Effects of Pharmaceuticals on Marine Fishes

Information on the biological impacts of these contaminants on marine fishes has been poorly investigated. In recent years, many studies have been carried out on freshwater ecosystems. In the next paragraph, we review the available literature analyzing the various effects of the most important pharmaceuticals in marine fishes, which are summarized in

Table 7. Biological effects of most common pharmaceuticals on marine fishes.

Therapeutical Class	Species	Molecule	Exposure Doses	Effects	Reference
Non-steroidal anti-inflammatory drugs	Oryzias latipes	Diclofenac	1.0 mg/L; 12–15 ng	Change in swimming; decrease in the number of eggs	[140]
	Trachyrincus scabrus	Diclofenac	100 μM	Interference with the metabolism of fish (CYP1A and CYP3A metabolism)	[141]
	Alocephalus rostratus	Diclofenac	100 μM	Inhibition of CE activity	[142]
	Solea solea	Ibuprofen	100 μM	Inhibition of the activity of the CYP3A4 enzyme	[143]
	Dicentrarchus labrax, Alocephalus rostratus, Cataetix laticeps	Ibuprofen	100 μM	Moderate inhibition of CE activity	[142]
	Solea solea, Dicentrarchus labrax, Alocephalus rostratus, Cataetix laticeps	Acetaminophen	100 μM	No change in CE activity	[142]
Antidepressant drugs	Solea solea, Dicentrarchus labrax, Alocephalus rostratus, Cataetix laticeps	Fluoxetine	100 μM	No change in CE activity	[142]
	Alocephalus rostratus, Mora moro	Fluoxetine	100 μM	Significant inhibition on BFCOD activity	[141]
	Opsanus beta	Fluoxetine	25 μg/g	Significant influence on branchial ureal excretion	[144]
Antibiotics	Pomatoschistus microps	Cefalexin	1.3–10 mg/L	Changes in feeding, behavior, biomarker responses, and predation performance	[145]
Cardiovascular drugs	Dicentrarchus labrax, Cataetix laticeps	Propanolol	100 μM	Inhibition of CE activity	[142]
Lipid regulators	Solea senegalensis	Gemfibrozil	1 mg/kg	Inhibition of antioxidant defences, induction of the activity of CYP-related and Phase II biotransformation enzymes	[142]
	Sparus aurata	Gemfibrozil	150 μg/L	Upregulated transcription of PPAR-related genes Increased cortisol levels	[146]
	Sparus aurata	Gemfibrozil	500 ng/L	Influence on the survival of larvae	[147]
	Carassius auratus	Gemfibrozil	1.5 μg/L	Reduced testosterone levels	[148]
	Sparus aurata	Gemfibrozil	1.5–15,000 μg/L	Increase in CAT and GR in the gills. Decrease in GPx and GR in the liver	[149]
	Dicentrarchus labrax, Trachyrincus scabrus, Alocephalus rostratus, Cataetix laticeps	Simvastatin	100 μM	Inhibition of CE activity	[142]
	Fundulus heteroclitus	Simvastatin	1.25 mg/L	Decrease in AChE levels	[150]
	Solea solea	Simvastatin	100 μM	Influence on CbE activity	[143]
	Dicentrarchus labrax, Solea solea, Trachyrincus scabrus, Alocephalus rostratus, Cataetix laticeps	Fenofibrate	100 μM	Inhibition of CE activity	[142]
	Solea solea	Fenofibrate	100 μM	Influence on CbE activity	[143]

and

Table 8. Biological effects of the common estrogen EE2 on marine fishes.

Molecule	Species	Exposure Doses	Effects	Reference
EE2	Sparus aurata	5 μg/g	Sperm motility and reduction in seminal liquid; changes in the levels of sex steroidal hormones and the gene expression profiles of various enzymes involved in the production of steroids; increases in the expression profile of the hepatic vtg gene; modulation in the expression of testicular protein and some hormone receptor genes in the gonads	[151]
EE2	Sparus aurata	50 and 500 ng/L	Decrease in the survival of fish larvae	[147]
EE2	Oryzyas melastigma	50–100 ng/L	Reduction in spawning and reproductive behavior	[152]
EE2	Oryzyas melastigma	90 ng/L	Growth retardation, reduction in embryos’ heart rate, decrease in the hatching rate, impaired larval locomotion, delays in the appearance of eye pigmentation	[153]
EE2 co-exposure to microplastics	Oryzyas melastigma	20–200 μg/L	Reductions in gonadosomatic and hepatosomatic growth and indices; increases in transcription levels of estrogen biomarker and estrogen receptor genes	[154]
EE2	Oryzyas melastigma	113 ng/L	Effects on immune function and reproduction of females	[155]
EE2	Cyprinodon variegatus	1000 ng/L	Alterations in the expression of estrogen genes	[156]
EE2	Cyprinodon variegatus	100 ng/L	Alterations in proteomic metabolism.	[157]
	Syngnathus scovelli	5 ng/L	Upregulation of estrogen biomarker genes	[158]
EE2	Dicentrarchus labrax	0.2–200 ng/L	Alterations in cytokine levels	[159]
EE2	Dicentrarchus labrax	0.5–50 nM	Alterations in the neuroendocrine gonadotropic system	[160]

.

3.1. Non-Steroidal Anti-Inflammatory Drugs

Non-steroidal anti-inflammatory drugs (NSAIDs) represent a significant class of pharmaceuticals that are commonly used for the treatment of pain and inflammation [161,162]. The aforementioned drugs include salicylates (aspirin), aryl alkanoic acids (diclofenac, indomethacin, nabumetone, sulindac), 2-aryl propionic acids or profens (ibuprofen, flurbiprofen, ketoprofen, naproxen), N-aryl anthranilic acids or fenamic acids (mefenamic acid, meclofenamic acid), pyrazolidine derivatives (phenylbutazone), oxicams (piroxicam, meloxicam), and sulfonamides (nimesulide) [1]. The chemical structures of the most commonly used NSAIDs are presented in Figure 1.

These classes of pharmaceuticals exhibit disparate structural, pharmacokinetic, and pharmacodynamic properties, yet they share a common biochemical effect. In fact, their mechanism of action is based on the inhibition of cyclooxygenase (COX) enzymes that catalyze the synthesis of prostaglandins and thromboxane from arachidonic acid [161,162]. Generally, cyclooxygenase enzymes consist of two isozymes: COX1 is responsible for maintaining baseline levels of prostaglandins, whereas COX2 induces the production of prostaglandins in response to stimulation at the site of inflammation. Consequently, classical NSAIDs inhibit both COX1 and COX2 equally, whereas new-generation NSAIDs are specifically targeted for the inducible isoform, COX2, which causes an inflammatory reaction [163].

Prostaglandins are produced in all tissues and cells, exhibiting a diverse molecular structure contingent upon the physiological role they fulfil. They are involved in a multitude of processes, including inflammation, regulation of blood flow in the kidney, pain, coagulation processes, and the synthesis of protective gastric mucosa [163]. In fish, prostaglandins are produced in several cells and tissues, including red blood cells, macrophages, and oocytes. A substantial body of research has been conducted on freshwater fish, demonstrating that NSAIDs can influence reproduction in these organisms. Examples of NSAIDs that have been shown to affect reproduction in fish include indomethacin and ibuprofen [164,165]. For instance, ibuprofen has demonstrated COX inhibition, which plays a pivotal role in fishes’ development and ovulation and maturation of oocytes. With augmented exposure, pairs have reproduced less frequently and spawned more eggs, resulting in a shift in breeding patterns [165]. Comparable effects have been documented in marine fish, although the effects of these classes of pharmaceuticals have been poorly investigated.

Among the non-steroidal anti-inflammatory drugs, diclofenac represents the most frequently detected compound in marine fish. Moreno-González et al. [166] evaluated the bioaccumulation of 20 significant pharmaceuticals in marine fish, including golden grey mullet (Liza aurata) and black goby (Gobius niger), considering their distribution throughout the Mar Menor Lagoon and their seasonal variations. With regard to non-steroidal anti-inflammatory drugs, diclofenac exhibited higher concentrations in the liver (2.2 ng/g) than in the muscle of these organisms. In general, all the psychiatric drugs in golden grey mullet were found to be preferentially present in spring, in contrast to the absence of seasonal differences in black goby. The bioaccumulation of pharmaceuticals was found to be lower in black goby than in golden grey mullet. This suggests that golden grey mullet may serve as a valuable indicator of the bioaccumulation of pharmaceuticals in marine environments.

Nassef et al. [140,167,168] demonstrated that the exposure of Japanese medaka Oryzias latipes to a concentration of 18 mg/L of diclofenac can result in the death of the organism within 24 h. Moreover, the behavior of O. latipes is also affected by exposure to low concentrations of diclofenac (1.0 mg/L), with a significant change in swimming [140]. A further study demonstrated that nanoinjections of diclofenac into medaka eggs had a significant impact on embryonic development. The study revealed that treatment with this analgesic at different doses (1.5 or 12 ng) significantly influenced embryonic development. In particular, it was reported that the number of embryos at 1 dpf (days post-fertilization) decreased considerably at a DCF dose of 12 ng/egg, whereas after 4 dpf, a reduction in the heart rate was observed compared with the controls [168].

Two studies [141,142] demonstrated that analgesics can lead to enzymatic inhibition in vitro in three marine fish species: the coastal marine fish Dicentrarchus labrax, the middle slope fish Trachyrincus scabrous, and the deep-sea fish Alocephalus rostratus and Cataetix laticeps. The species of fish examined include the European sea bass (Dicentrarchus labrax) and the sole (Solea solea), both of which are found in the middle slope, as well as the gadiform fish Trachyrincus scabrous, which is found in the middle slope, and the deep-sea fish Alocephalus rostratus and Cataetix laticeps. In particular, Solé and Sanchez-Hernandez [142] examined the sensitivity of carboxylesterase (CE) activity to diclofenac, acetaminophen, and ibuprofen. Their findings demonstrated that ibuprofen caused a moderate inhibition (28–40%) of CE activity in D. labrax, A. rostratus, and C. laticeps. The carboxylesterase activity of D. labrax, A. rostratus, and C. laticeps was found to be 40% lower than that of the controls. In contrast, diclofenac caused a significant inhibition of CE activity only in A. rostratus, while acetaminophen did not cause any change in CE activity. In contrast, Ribalta and Solé [141] analyzed the interference of diclofenac, which is also involved in xenobiotic and endogenous compounds, with the CYP1A and CYP3A metabolisms of T. scabrous. Moreover, the activity of benzyloxy-4-[trifluoromethyl]-coumarin-O-debenzyloxylase (BFCOD), a biomarker commonly used as an indicators of chemical exposure commonly used for monitoring CYP3A activity in fish, was demonstrably influenced by diclofenac in the coastal species D. labrax. Similar results were also obtained by Crespo and Solé [143], who investigated the biological effects of ibuprofen on juvenile S. solea using a multi-biomarker approach that integrated the impacts on different physiological pathways. It was reported that exposure to ibuprofen (100 µM) inhibited the activity of the CYP3A4 enzyme in the liver microsomal fraction of S. solea, whereas acetaminophen had no effect on this enzymatic activity.

3.2. Antibiotics

Antibiotics are considered to be the most significant pharmaceuticals discovered in the 20th century, resulting in a notable reduction in mortality and infectious disease worldwide [1]. These compounds are commonly used in both human and veterinary medicine to combat bacterial infections; however, they may also be used as promoters of animal growth [61]. In general, antibiotics are classified into several groups according to their different mechanisms of action, including the suppression of bacterial cell walls or protein synthesis and growth [10]. The most prevalent types of antibiotics are penicillins (e.g., penicillin and amoxicillin), which impede the synthesis of bacterial cell walls; tetracyclines (e.g., tetracycline), which bind to the ribosome and damage protein synthesis; and sulfonamides, which act to inhibit bacterial growth; quinolones (e.g., ciprofloxacin), and macrolides (e.g., erythromycin) [61,169].

In recent years, the impact of antibiotic pollution on aquatic ecosystems has been extensively studied in relation to the development of bacterial antibiotic resistance and the potential risks to human health. This is because the emergence and spread of antibiotic resistance in bacterial pathogens has led to a shift in the types of infections these pathogens can cause. [170,171,172]. Furthermore, the presence of other ubiquitous contaminants, such as heavy metals, can enhance co-selection processes and, consequently, antibiotic tolerance levels due to the co-regulation of resistance genes [36,172]. With regard to fish, the deleterious effects of antibiotics have been extensively investigated in freshwater ecosystems. However, in marine environments, the biological impacts of this class of pharmaceuticals on these organisms remain poorly understood. Nevertheless, it has been reported that antibiotic toxicity negatively influenced feeding, behavior, and biomarker responses in juvenile goby (Pomatoschistus microps) after exposure to cefalexin (at concentrations ranging from 1.3 to 10 mg/L) [145]. Furthermore, the study demonstrated that exposure to cefalexin at concentrations above 5 mg/L significantly reduced predation performance at 20 °C over 4 days, while lipid peroxidation levels increased (at 10 mg/L). At 25 °C, the toxicity of cefalexin increased, followed by a decrease in predation performance at 2.5 mg/L [145].

The chemical structures of common antibiotics are illustrated in Figure 2.

3.3. Antidepressant Drugs

Antidepressants represent a class of neuroactive pharmaceuticals related to psychiatric drugs, which also include antiepileptics and anxiolytics. These compounds are defined as selective serotonin reuptake inhibitors (SSRIs) and are used to treat a range of condition, including anxiety, clinical depression, social phobia, obsessive–compulsive disorders, panic disorders, and attention deficit disorders [1,61]. The occurrence of antidepressant residues and metabolites in aquatic environments has been confirmed through documented evidence. Silva et al. [173] reported values ranging from ng/L in WWTPs to a few µg/L in seawater. More recently, Duarte et al. [174] detected 33 neuroactive compounds in all fish brain and water samples and in 95% of the liver and fish muscle tissues analyzed. These emerging pollutants, including fluoxetine, citalopram, paroxetine, sertraline, and venlafaxine, exert their pharmaceutical action by inhibiting monoamine transporters and consequently the reuptake of the neurotransmitter serotonin (5-hydroxytryptamine, 5-HT) at the presynaptic neuronal membranes [175,176]. The chemical structures of common antidepressant drugs are presented in Figure 3.

The presence of serotonin (5-HT) in various tissues is a common feature among fish species, as evidenced by the presence of serotonin receptors [177]. In fish, serotonin functions as an important neurotransmitter, influencing a range of physiological processes, including behavior (aggression, appetite), endocrine, and reproductive parameters. Moreover, serotonin has been demonstrated to influence social hierarchy and feeding rank in certain species by exerting a profound influence on a multitude of processes, both at the individual and population levels. The influence is exerted through the alteration of social interactions and reproduction [169]. To date, only a limited number of studies have examined the biological effects of these pharmaceuticals on marine fishes. In contrast, a significant amount of scientific data has been reported for freshwater ecosystems. In particular, it has been demonstrated that antidepressants have deleterious effects on fishes’ physiology, reproduction, and behavior. In addition, there is evidence to suggest that they may act as endocrine disruptors in different species of fish [51,178,179,180].

For example, fluoxetine was demonstrated to affect the behavior of marine organisms. Such alterations include changes in growth rates, feeding habits, and predator–prey interactions [181]. In Ribalta and Solé [141], it was demonstrated that fluoxetine can inhibit cytochrome P450, which are enzymes involved in Phase I of xenobiotic and endogenous compounds. Moreover, fluoxetine has been demonstrated to affect the behavior of marine fishes by reducing locomotory activity after 32 h of exposure at an effective concentration of 155 µg/L [182]. Conversely, the administration of fluoxetine to the Gulf toadfish Opsanus beta via the intraperitoneal route was demonstrated to influence branchial urea excretion and intestinal osmoregulation at a high concentration., resulting in a stress response and a subsequent elevation of plasma cortisol levels [144]. In Solé and Sanchez-Hernandez [142], the antidepressant fluoxetine had no effect on the CE activity of marine fishes (S. solea, D. labrax, T. scabrux, A. rostratus, C. laticeps, and Siganus cataliculatus), which is a fundamental mechanism involved in pesticide detoxification of fish [183].

3.4. Cardiovascular Drugs

The pharmacological management of cardiovascular disorders typically entails the use of drugs that act on the cardiovascular system, including β-blockers (such as atenolol, propranolol, metoprolol), angiotensin-converting enzyme (ACE) inhibitors, angiotensin II antagonists (includingsartans and valsartan), and calcium channel blockers (such as diltiazem). The chemical structures of common cardiovascular drugs are shown in Figure 4.

It has been demonstrated that pharmaceuticals of the aforementioned classes are frequently detected in aquatic environments, particularly in the effluent of wastewater treatment plants and surface waters. Concentrations of these pharmaceuticals at ng/L levels and up to 10 µg/L have been observed [184,185,186,187,188,189,190,191,192,193].

β-blockers represent a category of cardiovascular pharmaceuticals that are the most frequently prescribed for the treatment of a range of cardiac conditions, including angina, high blood pressure, and glaucoma. Their mechanism of action is based on specific binding to adrenoreceptors, competing with β-adrenergic agonists and leading to a decrease in heart rate, cardiac output, and cardiac muscles’ contractibility. A number of research studies have documented the physiological effects of β-blocker pharmaceuticals on freshwater fish species. These effects have been observed in tissues and organs outside the heart, including branchial vascular tissue, the gills, the liver, erythrocytes, the brain, and skeletal muscle [181,194,195]. It has been demonstrated that three subclasses of β-blockers (β1, β2, and β3) have different levels of potency and efficacy. It has been demonstrated that these receptors are involved in numerous physiological processes in fishes. These include the regulation of cardiovascular function, growth, and metabolism [169,196].

Only recently have the biological effects of cardiovascular drugs been studied in marine fishes. In particular, the effects of β-blockers have been evaluated in vitro. Propranolol exposure resulted in a decrease in the enzymatic activities of carboxylesterase (CbE) and benzyloxy-4-trifluoromethylcoumarin [142,143].

3.5. Lipid Regulators

The category of lipid regulators includes two groups of agents that lower cholesterol and triglycerides in the blood plasma: statins and fibrates. These agents play an important therapeutic role in the treatment of hyperlipidemia. Statins, such as simvastatin and atorvastatin, are known to inhibit the activity of the enzyme HMG-CoA (3-hydroxymethylglutaryl coenzyme A reductase), which is responsible for feedback control of cholesterol synthesis [61]. The result of their activity is a decreased intracellular cholesterol concentration and an overexpression of low-density lipoprotein (LDL) receptors on hepatocyte membranes. This results in the clearance of circulating LDL cholesterol [61]. The chemical structures of some of the common lipid regulators are shown in Figure 5.

Peroxisomal proliferators (PPs) such as fibrates, including bezafibrate and gemfibrozil, exert their therapeutic effect by inducing changes in the expression of genes involved in lipoprotein metabolism. These changes include increased expression of enzymes such as acyl-CoA oxidase (AOX), which facilitates peroxisomal β-oxidation of fatty acids [197]. PPs specifically act by binding to and activating the peroxisomal proliferator-activated receptor alpha (PPAR-alpha) transcriptional factor, which then binds to specific response elements (PPREs) in the promoter region of PP-sensitive genes. This ultimately results in the removal of fatty acids and cholesterol from the blood [61,198].

With regard to the adverse effects of antilipidemic drugs on marine organisms, only a limited number of studies have been conducted to assess the impact of these pharmaceuticals on the development and reproduction of invertebrates. In marine fishes, biological responses have been evaluated through molecular and biochemical changes. Indeed, it has been demonstrated that fishes possess a peroxisomal β oxidation system, and these organisms respond to PPAR agonists by increasing AOX activity. This indicates that the PPAR pathway represents a fundamental factor in mediating enzymatic responses to fibrates in fish [199,200].

In marine fish, studies on the toxicity of lipids have focused on the biological effects of the fibrate gemfibrozil, the most prescribed pharmaceutical in human medicine [201]. It has been demonstrated that this compound induces alterations in the balance of sex hormones and embryonic malabsorption syndrome in zebrafish [148,202]. Teles et al. [146] demonstrated that exposure to gemfibrozil (150 g/L) upregulated peroxisome proliferator-activated receptors (PPAR) and transcription of the related genes in juvenile Sparus aurata, resulting in an increase in cortisol levels as a stress-related effect at a concentration of 1.5 mg/L. Some studies indicated that lipid regulators can cause variations in fishes’ metabolism. Recently, Barreto et al. [149] demonstrated that exposure of S. aurata to 1.5–15,000 g/L of gemfibrozil resulted in an increase in catalase (CAT) and glutathione reductase (GR) in the gills and a decrease in glutathione peroxidase (GPx) and GR in the liver.

In Solé et al. [142], the administration of gemfibrozil at 1 mg/kg body weight in Solea senegalensis resulted in the activation of cytochrome P450-related and Phase II (UDPGT) biotransformation enzymes, while simultaneously inhibiting the antioxidant defenses. Conversely, in Mimeault et al. [148], gemfibrozil reduced serum testosterone levels in the goldfish Carassius auratus at 1.5 g/L. In contrast, Ribalta and Solé [141] reported that gemfibrozil did not cause significant interference with ER metabolism in any fish species studied. This highlights that the interference with the CYP3A is species-dependent, with higher levels observed in the perciform D. labrax (26%). More recently, the effects of gemfibrozil on marine fish were also investigated by Capolupo et al. [147], who demonstrated that this compound negatively affected the survival of the larvae of the seabream S. aurata following exposure to 500 ng/L of gemfibrozil. Furthermore, other fibrates, such as simvastatin and fenofibrate, have been demonstrated to inhibit CbE activity in coastal marine fish (D. labrax), middle-slope fish (T. scabrus), deep-sea fish (A. rostratus and C. laticeps) [142], and S. solea [143]. In addition, simvastatin exposure has been demonstrated to result in a reduction in AChE levels in estuarine Fundulus heteroclitus (1.25 mg/L and LC₅₀ of 2.68 mg/L) [150].

3.6. Steroid Hormones and Estrogens

A number of studies have provided evidence to demonstrate the adverse effects of steroidal hormones in the environment, highlighting the deleterious impacts on non-target marine species at concentrations lower than those of bioactive substances [147,203,204]. The assessment of steroid hormones commenced in the early 1970s, and this led to the first studies of pharmaceuticals in aquatic environments. Both natural and synthetic steroids are widely used in human and veterinary fields, with large quantities discharged from livestock and aquaculture [204,205,206]. The endocrine system represents the principal target of steroid hormones due to its role in maintaining the homeostasis of organisms, and their development, behavior, and reproduction [204]. The principal effects are attributable to estrogens, particularly 17-ethynilestradiol (EE2) (Figure 6) It is considered to be the primary component of oral contraceptives and the synthetic drug with the highest potential to disrupt the endocrine system [203].

As reported in some studies on mammals and marine organisms, physiological estrogens derived from cholesterol [207,208] and the mechanism of direct genomic signaling is based on modulation by nuclear estrogen receptors (ERs), which play a role as ligand-activated factors [209,210]. Following the binding of estrogens in the cytoplasm, ERs undergo a conformational change, including receptor dimerization. This complex then migrates into the nucleus, where it binds to chromatin through the estrogen response element (ERE), stimulating the transcription of the target genes. It has been demonstrated that estrogens can modulate the transcription of other genes through indirect genomic signaling after the interaction of ER complexes with other transcription factors and response elements, inducing the activation or suppression of the target genes’ expression [4]. In addition to genomic signaling, the non-genomic mechanism represents another important pathway through which estrogens can determine cellular effects. These mechanisms are based on the interaction of membrane-bound ERs, which then activates signaling cascades such as the Ras/Raf/MAPK cascade, the phosphatidylinositol 3-kinase/Akt kinase cascade, and the cAMP/PK A signaling pathway [4], which regulates intracellular concentrations of Ca²⁺ and nitric oxide, and the expression of ERE and transcription factor regulator genes. To date, several studies have been conducted in teleost fishes to illustrate the negative effects of estrogens on the endocrine system, reproduction, behavior, development, metabolism, and cytokine expression [151,152,153,155,156,157,158,159,160]. In particular, three ERs have been described (ER1, ER1, and ER2), which exhibit some similarities in the activity of EE2 between mammals and fishes in the molecular pathway linked to the signaling system [211]. It has been observed that the production of vitellogenin, a protein precursor induced by estrogen, is increased in fishes. This protein is produced only by adult females, but it has also been identified in juvenile and male fishes exposed to estrogen-like substances. The synthesis of vitellogenin is linked to several events. Firstly, the activation of the ER signaling pathway by estrogens leads to an increase in the expression of the vitellogenin gene. Secondly, the protein is mobilized and matured in the endoplasmic reticulum and Golgi apparatus. Finally, it is incorporated into secretory vesicles and released into the circulatory system [4].

It has been documented that the global human population releases approximately 30,000 kg of natural steroid estrogens into the environment annually, in addition to 700 kg of synthetic estrogen, which arises exclusively from the use of birth control pills [212]. The environmental release of estrogenic EPP through wastewater and effluent is a frequently observed phenomenon, as evidenced by studies conducted in the UK and in Europe. This has resulted in deleterious effects on fish.

For instance, the use of synthetic estrogens in contraceptives has been associated with the feminization of male fish, which has resulted in skewed sex ratios and reproductive impairments [213].

The bioaccumulation of estrogenic EPP has been observed to be associated with the phenomenon of the feminization of male fish [5].

The result of this process is intersex, which is the presence of male and female reproductive tissues. A higher concentration of the estrogenic compound was observed to correspond with a greater occurrence and intensity of feminization [214]. Furthermore, intersex was rarely observed in fish less than 3 years of age, in contrast to more advanced age classes, where it was more common. This indicates that the expression of this condition is progressive [5].

A detrimental consequence observed in fishes is the feminization of males following exposure to EE2, which is then accompanied by the development of ovotestes in species that lay eggs. This phenomenon is associated with the presence of oocytes in the male gonads, impaired spermatogenesis, and reductions in sperm motility and sperm counts. In addition, there is evidence that exposure to EE2 affects sperm motility and sperm counts [203,204,215,216]. Furthermore, exposure to EE2 has been linked to an increased frequency of cancers, which is thought to be due to the role of EE2 as a promoter of the formation of hepatic tumors through a reduction in the ability to repair DNA adducts [204,216].

The adverse physiological impacts of estrogen have also been evaluated in behavioral studies, which have demonstrated the consequences at population levels. In particular, it has been reported that in sand gobies Pomatoschistus minutus, exposure to EE2 at a concentration of 41 ng/L for 31 days caused altered reproductive behavior, influencing the movements of attraction of females’ attention and males’ parental care in developing eggs [217]. Conversely, it has been observed that exposure to low concentrations of EE2 (1 ng/L, 10 days) induced negative effects on the expression of secondary sexual traits and the mating dynamics of the Gulf pipefish Syngnathus scovelli [218]. Furthermore, a correlation has been demonstrated between microplastics and the estrogenic effects of EE2 in O. melastigma, indicating an increase in the estrogenic impacts of EE2 in marine fishes [154].

4. Conclusions and Future Perspectives

The environmental impact of pharmaceuticals has recently been identified as a significant global threat and a key research area in marine science. The persistence of these emerging pollutants and their degradation products is largely attributed to their unregulated and continuous spread in the environment, which has been linked to a wide range of detrimental effects. These effects can be observed at the organism level, including the onset of antibiotic resistance phenomena. Currently, the available information on the effects of pharmaceuticals on non-target organisms is still heterogeneous and fragmented, with a significant focus on the effects of NSAIDs and psychiatric drugs compared with cardiovascular drugs and antibiotics.

With regard to marine fish, further scientific knowledge is required in order to enable a more detailed investigation of the impacts of certain therapeutic compounds, including antibiotics, cardiovascular drugs, lipid regulators, and hormones, on these organisms in terms of toxicity, bioaccumulation, metabolic processes, and trophic transfer. Furthermore, the long-term effects of exposure to environmentally relevant concentrations must be evaluated. This review has highlighted the complexity of the responsiveness and variations in the regulatory mechanisms of marine fishes to pharmaceuticals, which may be further challenged by the simultaneous presence of other contaminants such as microplastics and potential environmental stressors (i.e., increased temperatures). Therefore, the reviewed data must be interpreted with a certain degree of caution due to the high variability in the reported biological results, doses and modes of exposure, typologies of pharmaceuticals analyzed, and characteristics of non-targeted organisms.

Future challenges and including monitoring studies should be carried out with the aim of developing new strategies for evaluating the impacts and acute effects of mixtures of pharmaceuticals and other pollutants on marine fishes due to the lack of data. Moreover, scientific research should enhance the development of a “green pharmacy” based on more sustainable pathways for the design, production, and degradation of synthetic drugs in wastewater. This approach has the potential to reduce the impacts on the marine environment and related organisms.