Impact of Micro- and Nanoplastics on Aquatic Organisms: History
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Contributor:
Eugenio Geremia
,
Maria Teresa Muscari Tomajoli
,
Carola Murano
,
Adriana Petito
,
Gianluca Fasciolo

Anthropogenic activities on the aquatic ecosystem have direct and indirect effects that negatively affect biodiversity both in freshwater and sea. Among the main threats to aquatic ecosystems and biodiversity is climate change. Climate change can impact marine and freshwater ecosystems in several ways: ocean warming; sea level rise; the loss of sea ice; a decrease in pH due to increased ocean surface acidity; an increased risk of diseases in marine biota; the loss of habitats such as Coral Reef; pollution; nutrient enrichment; hydrological modifications; the spread of invasive species; and increasing levels of UV light. The production of plastic and its release into the environment is contributing to the loss of biodiversity and is becoming a serious threat to animal redox homeostasis and, subsequently, for human health.

  • microplastics
  • nanoplastics
  • toxicity

1. Pollution and the Environment

The increase in the world population and, consequently, human activities, such as agricultural intensification, industrial development, and urbanization, have led to a sharp increase in waste production and environmental pollution [1][2]. Coastal, marine, and fluvial environments are constantly under pressure from anthropogenic release into the environment of heavy metals, hydrocarbons, pesticides, persistent organic pollutants (POPs), organophosphorus flame retardants (OPFRs), pharmaceuticals and personal care products (PPCPs), products of metallurgical processes, nanoparticles, plastics and microplastics, etc. [3][4][5][6][7][8]. Organisms living in the aquatic environment, such as plants and animals, can transfer pollutants to humans through bioaccumulation and trophic transfer, as well as exert harmful effects on them following their exposure [5][7]. In recent decades, there has been a rise in the emission of nanoparticles and heavy metals into the environment with the intensification of industrialization [9][10][11]. Nanoparticles are a group of materials that have a size between 1 and 1000 nm in diameter [12][13][14]. The nanoparticles are widespread in aquatic ecosystems through aerial deposition, effluent discharge, dumping, and run-off [11]. They can be divided into two groups: natural (i.e., desert dusts, aerosols, emissions from volcanic activities, etc.) and anthropogenic (i.e., metal oxides, drug production, burning fossil fuels, vehicle exhaust release, mining demolition emissions, etc.) [11][12][13]. One of the most widespread nanoparticles in the aquatic environment is titanium dioxide (TiO2), which the European Union reports as one of the main pollutants of surface waters, with a value of 2.2 μgL−1 [12]. Lovern and Klaper, showed an increase in the mortality rate of D. magna due to exposure to TiO2 [14]. Heavy metals also pose a serious threat to the aquatic environment. Twenty-three metals and metalloids out of seventy present in the environment are identified as heavy metals/trace metals, some of which are considered dangerous [9]. Heavy metals can be divided into two categories: essential and non-essential. The non-essential metals (aluminum (Al), cadmium (Cd), mercury (Hg), tin (Sn), lead (Pb), etc.) are elements that have no known biological function and can be considered toxic at high concentrations. Essential metals (copper (Cu), zinc (Zn), chromium (Cr), nickel (Ni), cobalt (Co), molybdenum (Mo), and iron (Fe)) play a specific role in the body’s metabolism and can be toxic both under metabolic deficiency conditions and at high concentrations [9][10][15]. Contaminants/heavy metals can have a negative impact on the physiology of aquatic organisms; fish especially are very sensitive to such changes [15]. The toxic effects of heavy metals on fish are numerous [16]. They enter the fish body via the gills, digestive tract, and body surface and accumulate in the liver, kidney, muscle, intestine, skin, and bones [16]. It is shown that in fish heavy metals can alter several physiological and biochemical processes, growth rates, mortality, and reproduction. Moreover, they can cause serious problems in the body by producing reactive oxygen species (ROS), which cause oxidative stress and DNA damage [17][18][19].
Another issue that has plagued the environment in recent decades is the spread of pharmaceuticals and personal care products (PPCPs) in aquatic environments. This group of pollutants includes several substances from different sources, mainly anti-inflammatory drugs, as well as antibiotics (antimicrobials or antibacterials), antiepileptics, and personal care products [20]. The increase in the consumption of PPCPs is caused by the average use in the use human population and the upward demand for animal protein [21][22][23][24]. These substances find their way into the environment, where they can negatively impact organisms in aquatic ecosystems. Antibiotics are the most widely used pharmaceuticals due to human and veterinary applications [25]. Environmental antibiotics are derived from different sources, such as wastewater treatment plants, hospitals, water from urban canals, agriculture (aquaculture, husbandry), and the pharmaceutical industry [26][27]. Recently, the use of antibiotics has increased with the rise in aquaculture; in fact, their use allows for the reduction in possible threats from microorganisms. However, on the other hand, antibiotics can promote the resistance of bacteria in aquaculture and lead to resistance even in wild fish populations [28]. They are continually discharged into the aquatic environment and are bioavailable for animals, crops, and aquatic plants [26][27][28]. Primary producers, like microalgae and decomposers, are more sensitive than crustaceans and fish to antibiotic agents. Consequently, antibiotics could negatively impact microbial processes and lead to variations in biogeochemical cycling and aquatic ecosystems [4][26].
Among the toxic pollutants with the greatest impact on the aquatic environment, people cannot fail to include pesticides [29][30][31][32]. Pesticides are important for the protection of crops from infection by many microorganisms and for eradicating pests and organisms, such as mice, rats, ticks, and mosquitoes, that are dangerous for the environment and the habitat of humans [30][32][33]. Among these, the best known and most discussed is DDT (1,1,1-trichloro-2,2-bis(4-chlorophenyl) ethane) [30][33]. However, due to its high toxicity, it has been banned in many Western countries since the 1970s [33]. Due to its ecotoxic effects, bioaccumulation, and environmental persistence, it was also considered by the international agreement of the Stockholm Convention 2001 (World Health Organization, 2011) as one of the world’s 12 restricted persistent organic pollutants [34]. Nevertheless, the World Health Organization controlling vector-borne diseases allowed it in certain tropical countries in 2006 [30]. DDT can land many issues for aquatic organisms. It is so hydrophobic that it is absorbed into organic sediment particles, where it can persist for many years. In the aquatic environment, it can accumulate in benthic organisms and subsequently be transferred to higher trophic levels through biomagnification [30]. Owens e Baer, 2000, revealed a dose-dependent increase in the lethality of the Japanese medaka (Oryzias latipes) exposed to DDT [35], while Ton et al., 2006, showed the teratogenic effects on zebrafish (Danio Rerio) larvae exposed to the organochlorine DDT [36].

2. The Impact of Micro- and Nanoplastics on the Aquatic Environment

Pollution by human activities (agriculture, industrialization and urbanization) can take many forms, including air, water and soil pollution, electromagnetic pollution (radiation waste), and even noise pollution. These may contribute to health issues and decline the quality of life [37]. Water pollution by agricultural, municipal, and industrial sources have become a major concern for the well-being of humanity and biodiversity [38]. Anthropogenic activities on the aquatic ecosystem have direct and indirect effects that negatively affect biodiversity both in freshwater and sea [39][40][41]. Among the main threats to aquatic ecosystems and biodiversity is climate change. Climate change can impact marine and freshwater ecosystems in several ways: ocean warming; sea level rise; the loss of sea ice; a decrease in pH due to increased ocean surface acidity; an increased risk of diseases in marine biota; the loss of habitats such as Coral Reef; pollution; nutrient enrichment; hydrological modifications; the spread of invasive species; and increasing levels of UV light [41]. In addition to climate change, the production of plastic and its release into the environment is contributing to the loss of biodiversity and is becoming a serious threat to animal redox homeostasis and, subsequently, for human health [5].

2.1. MPs/NPs’ Effects on Phytoplankton

Normally, micro- and nanoplastics, due to their small size and low density compared to water, tend to float in the water column. When microplastics enter the aquatic environment, they are rapidly colonized by microorganisms and algae, typically within hours [42]. It has been highlighted that a microbial biofilm instantly colonizes plastics when exposed to the environment. It is estimated that between 1000 and 15,000 metric tons of microbial biomass are harbored on plastic marine debris [43]. However, the biomass accumulation due to biofouling can increase its density and cause it to sink [44]. Microalgae also can adhere to particles such as MPs/NPs. Many microalgae secrete polysaccharides if stressed due to decreased nutrients and light. Exopolysaccharides may coagulate to form sticky particles named transparent exopolymer particles (TEPs). TEP favors microalgae aggregation; these aggregates are the primary vehicle for transporting phytoplankton and debris towards the seabed [45]. Another sedimentation pathway consists of small MPs ingested by zooplankton and deposited within fecal pellets [46].
MPs and NPs show effects on each trophic level. Microalgae are the first level in the food chain and are indispensable for the marine ecosystem equilibrium. MPs seem to affect the well-being of microalgae. As has been reported, different sizes and concentrations of MPs and NPs can inhibit growth, reduce chlorophyll and photosynthesis, induce oxidative stress, cause changes in morphology, and promote the production of heteroaggregates [47][48].
Janak Raj Khatiwada (2023), evaluated the effect of PET microplastic (25 mgL−1, 50 mgL−1, 100 mgL−1 and 200 mgL−1) of Scenedesmus sp. for 24 days. PET showed the highest growth inhibition effect at 200 mg/L. Moreover, compared to the control group, microalgae exposed to microplastics showed a lower chlorophyll content, possibly due to blocking the pores of MPs for cellular respiration [49]. Cai Zhang (2017) carried out tests on Skeletonema costatum, demonstrating that microplastic (Micro-PVC, average diameter 1 μm) significantly inhibited the growth of microalgae and chlorophyll at 96 h under 50 mg/L treatment [50]. Microplastics may also affect microalgae lipid and fatty acid composition. Guschina et al. (2020) assessed the exposure of Chlorella sorokiniana to polystyrene microplastics (<70 μm, 60 mgL−1) for four weeks. The study showed that microplastic could alter the concentrations of essential fatty acids, which are necessary for algae’s lipid synthesis. Such changes could also have repercussions for food quality, growth, and stress resistance in primary consumers, and could affect potential propagation through trophic transfer [51]. On the other hand, microalgae can also produce effects on microplastics. Among these, it is possible to enumerate the different alterations of MP properties; in particular, their adsorption seems to be enhanced [42]. Wang et al., (2020) showed that the development of a biofilm altered the adsorption of copper Cu(II) and tetracycline (TC) by PE microplastic [52]. The biofilm can also camouflage the plastic particles and promote their ingestion by grazers, such as zooplanktonic organisms [42].

2.2. MPs/NPs in Sediments and Benthos

When micro- and nano plastics lose their ability to float, they can fall and settle on the sea bottom. This severe issue can be observed primarily in coastal shallow-water regions. However, these pollutants are not limited to the coastline; they have also been identified in deep-sea sediments with concentrations of up to 2000 particles m−2 at a depth of 5000 m [44][53]. Finally, these small-sized materials will find their pathway to the benthos, affecting some marine organisms. Several studies in the laboratory have investigated the ingestion of micro- and nanoplastics by benthic organisms.
Cole et al. (2020) have compared the toxicity of microplastics and nanoplastics on Mytilus edulis for 24 h or 7 days. Mussels were exposed to 500 ng mL−1 of 20 μm polystyrene microplastics, 10 × 30 μm polyamide microfibers, or 50 nm polystyrene nanoplastics. After 24 hours of exposure, there was a rise in SOD activity, but after 7 days, it returned to normal levels without negatively impacting health. Particle size, however, can influence sub-lethal toxicity because exposure to nanoscopic plastic raises the proportion of phagocytic hemocytes (indicating a heightened immune response) and leads to a significant increase in micronuclei formation [54]. The mussels are very important in these studies because they represent the prey of many intertidal species and serve as a food source for humans [55].
Sussarelli et al. (2016) assessed the impact of the polystyrene microspheres (micro-PS) on the physiology of the Pacific oyster (Crassostrea gigas). The organisms were exposed to virgin micro-PS (2 and 6 µm in diameter; 0.023 mgL−1) for 2 months during a reproductive cycle. Ingestion preference was shown for the 6 µm particles over 2 µm particles. After 2 months, the histological analysis only detected micro-PS particles in the stomach and intestine. Hyalinocytes and granulocytes in exposed oysters were larger than in controls. The total number of oocytes and the oocyte diameter were significantly lower in exposed females than in controls, while exposed males had a considerably lower sperm velocity. Finally, in progeny issued from exposed genitors than in progeny issued from control genitors, the larval growth and size were significantly slower [56].
Murray and Cowie (2011) assessed the capability of the Nephrops norvegicus, an ecologically and commercially important crustacean, to consume plastics. They evaluated the stomach microplastic content of shrimp through trawls in the north Clyde Sea area. Four tanks were set up, and the organisms were fed fish seeded with approximately 5 mm strands of blue polypropylene rope (ten filaments for one cm3 of fish). Fish without plastic seeds were provided to the animals in the control group. Nephrops were killed after 24 h, and new animals were placed in each tank. This experiment was repeated five times over a 2-week period. Nephrops with no plastic in their stomachs had a larger median carapace length than those with plastic. This is presumably due to the capacity of the larger animals to sort plastics from food or excreting plastic once ingested. This study demonstrates that N. norvegicus can passively accumulate plastic in the stomach via during feeding. There could be potential implications for human health because the Nephrops is a commercially fished species [57].

2.3. Trophic Levels: From Zooplankton to Main Fishes Consumed by Humans

Considering the enormous variety of fish and the environments in which they live [58], humans make use of fish in many ecosystem services [59][60]. Precisely in this regard, it must be considered that human activities are impacting fish biology and their redox homeostasis, on whose alteration micro- and nanoplastics play an increasingly decisive role.

2.3.1. MPs/NPs’ Effects on Zooplankton

Zooplankton is the second food chain level and represents a key trophic link in pelagic food webs. The role of zooplankton is vital in the aquatic environment, as they are primary consumers and include the juvenile stages of numerous commercially important species (e.g., the meroplankton) [61]. Cole et al. (2013) reported that the ingestion of polystyrene beads, already starts with the zooplankton, namely the copepod Centropages typicus, showing the relevance of the MPs in significantly reducing the consumption of algae. This study implies that marine microplastic debris can negatively impact health and zooplankton function [61]. Ziajahromi et al. (2017) investigated the effects of microplastics on the freshwater zooplankton Ceriodaphnia dubia. The acute (48 h) and chronic (8 d) effects of microplastic polyester fibers and 1–4 μm PE beads on zooplankton have been examined. This study demonstrated the microplastic fibers caused a 50% reduction in reproductive capacity at concentrations approximately six times higher than environmental concentrations. Unlike other studies, this study did not observe any ingested fibers. However, malformations have been observed in the carapace of organisms exposed to polyester fibers. However, malformations have been observed in the carapace of organisms exposed to polyester fibers. This demonstrates that the negative effects of microplastic fibers on exposed aquatic organisms can also include external physical damage and not only those resulting from ingestion [62]. Rehse et al. (2016) investigated the effects of 96 h exposure of PE-particles (size 1 μm) on Daphnia magna. The study showed that the particles are ingested and cause immobilization. The immobilization of daphnids increases with dose and time. EC50 was 57.43 mg L−1 [63].
Another species utilized to study the effects of MPs/NPs is the marine copepod Calanus helgolandicus. This organism is a keystone species that can constitute up to 90% of mesozooplankton biomass within marine waters throughout Europe and the northeast Atlantic. Calanus helgolandicus has a large size and high lipid content; its abundance makes it a vital prey species for the larvae of a number of commercially important fish. Cole et al. (2015) exposed Calanus helgolandicus to 20 μm polystyrene spheres and cultured algae for 24 h to evaluate ingestion rate and for 9 days to evaluate reproductive function. The results demonstrate that microplastics can alter the ingestion of algae and reduce the size of the eggs, probably due to a reduction in the ingested carbon biomass. Moreover, reducing ingested carbon could cause an energy depletion and the consequent death of the organism [64].
The effects of microplastics are not always so evident, as in the case of exposing the MPs to Pacific oyster larvae (Crassostrea gigas). The exposure has not shown any significant effects on the development or feeding capacity of the larvae. This could be because of the oyster’s more simplistic intestinal tract [65].
Zheng et al. (2020) suggested an interesting point of view. MPs ingested by copepods, the crustacean zooplankton that is the main prey of fish, are likely to be carried up the food chain, while those ingested by jellyfish, which have fewer predators than copepods, are more likely to be discharged into the marine environment and circulate in seawater or seabed sediments. In this regard, the distribution of MPs, in addition to depending on the chemical-physical characteristics already discussed, also depends on the different groups of zooplankton [66].

2.3.2. MPs/NPs’ Effects on Others Aquatic Trophic Levels

The effects of these pollutants on fish vary depending on the dose, the target organisms, and the interactions between the pollutants [67][68]. Nanoplastics have been reported to have a higher impact than microplastics [69][70]. However, this is not always true. As reported by Jiang et al. (2020), there are conditions in which microplastics can have a greater impact than nanoplastics [71]. These differences can be attributed to multiple factors, including the capacity of ingestion, the possibility of entering tissues and cells, and the type and shape of the polymer that can affect the particles’ interaction with other pollutants and chemical compounds [72][73][74].
Microplastics may not always have “negative” effects. As already stated, based on the shape and size of the polymer, microplastics can interact with other chemical pollutants. This capacity, which also depends on the dose, could mitigate the effect of chemical pollutants (antagonistic effect). If, on the one hand, MPs can reduce pollutants’ effects, on the other, they can act synergistically by enhancing its effects (agonist effect), which in the latter case is potentially more toxic than nanoplastics [75][76][77]. Environmental factors can also dictate the dose-dependent impacts of microplastics. For example, phenomena such as storms and cyclones can increase their concentration and thus induce an increment in dose in specific areas, leading to an increase in the absorption by aquatic organisms [69][78].
Neves et al. (2015) reported the presence of microplastics in 63.5% of benthic fish and 36.5% of pelagic fish species, with a total of 73 microplastics identified from fish stomach contents [79]. A study on rainbowfish (Melanotaenia fluviatilis) exposed to microbeads adsorbed with polybrominated diphenyl ethers (PBDEs) was monitored for 0, 21, 42, and 63 days. Exposed fish accumulated high concentrations of PBDEs (ca.115 pg g−1 ww per day) in tissue after ingestion [80]. Redondo-Hasselerharm et al. (2018) report that in Gammarus pulex exposed to sediment containing MPs (Micro-PS 20–500 μm for 28 d), the growth is reduced [81], as well as in Gammarus fossarum [82]. A study conducted on juvenile European perch (Perca fluviatilis) exposed to polystyrene microplastic particles (90–150 μm) over six months showed that the animals ingested and accumulated the polystyrene microplastics which resulted in reduced growth; delayed hatching; and impaired performance, nutrition, and behavior [83].
MPs and NPs also influence the olfactory senses, which increases susceptibility to being killed by predators [84][85]. The predator–prey relationship gradually increases the toxicity of these substances as they are transferred from organism to organism, accumulating [40][86][87][88]. About 18% of the main predators of the central Mediterranean, Swordfish (Xiphias gladius), bluefin tuna (Thunnus thymus), and albacore (Thunnus alalunga) ingested micro-, meso-, and macroplastic debris ranging in size from <5 mm to 5–25 mm to 25 mm, respectively, [89][90] found microplastics in 36.5% of the gastrointestinal tracts of pelagic and demersal fish. Tuna is the main consumed fish; in fact, much attention is being paid to the sustainability of its fishing by FAO and the UN [91]. Dias-Basantes et al. (2022) showed that canned tuna can provide an average of 692 ± 120 MPs/100 g in brine-soaked tuna and 442 ± 84 MPs/100 g in oil-soaked tuna, values that significantly exceed those reported in research on canned fish [92]. Di Giacinto et al. (2023) showed the presence of MPs (size < 10 μm), polymers (PET, polycarbonate (PC)), and additives (Bisphenol A (BPA) and p-phthalic acid (PTA)) in the muscle Thunnus Thynnus and Xiphias Gladius fished in the Mediterranean Sea [93]. Abihssira-García et al. (2020) report that MPs were also found in Atlantic Salmon (Salmo salar), the second main fish consumed by humans. They show that Atlantic salmon immune cells from blood, distal intestine, and head kidney can phagocytose MPs (1–5 μm) even at relatively low concentrations (low 0.05 mgL−1; medium 5 mgL−1; high 50 mgL−1 for 1, 24, 48 and 72 h) and that their mortality is affected by the time exposure and the microplastic type [94].
Another aspect concerning the trophic levels that should be considered is that the ingestion and accumulation of microplastics take place before maturity, impacting the number of available organisms. A study conducted on Artemia nauplii (the early stage of development that occurs after the cysts hatch), which were subjected to a high concentrations of microplastics (1.2 × 106 particles per 20,000 nauplii), showed that they had ingested and accumulated microplastic particles ranging in size from 1 to 20 μm, in high concentrations; these particles were subsequently transferred to the zebrafish that fed on the nauplii [95]. Although not all particles would have been transferred from the nauplii to the zebra, as they were partially excreted, some were retained within the epithelial cells and intestinal villi. Another study on the absorption and effect of microplastics on zebrafish showed that most of the plastic particles (5 μm in diameter) had accumulated in the gills, intestines, and liver. Those with a larger diameter (20 μm) could accumulate only in the intestine and gills. Thus, the smaller particles (5 μm) are more toxic than the bigger ones (20 μm), causing inflammation and the accumulation of lipids in the fish liver, inducing oxidative stress, and altering the metabolic profiles of the fish liver, disturbing lipid and energy metabolism [96]. Accordingly, Jeong et al. (2016) showed that exposure to plastic beads (0.05, 0.5, 6 μm) in Brachionus koreanus was associated with increased oxidative stress and decreased growth rate, fecundity, lifespan, reproduction time, and body size, suggesting that the toxicities of MPs and NPs are size-dependent and that smaller plastics are more toxic than bigger ones [97]. This size dependency is probably correlated with the ability of smaller molecules to enter cellular compartments. Manabe et al. (2011) reported that smaller NPs were more easily ingested by medaka (Oryzias latipes) than the bigger ones and that NPs were excreted more slowly than the microplastics [98]. NPs negatively affect the liver health of medaka fish, leading to significant alterations in digestion, innate immune, and antioxidant-enzyme-based liver tissue damage [99] and inflammation [100]. However, microplastics are capable of damaging cellular organelles. In line with that, Felix et al. (2023) find that MPs in zebrafish can affect the center of the energy balance: mitochondria. After 21 days of exposure to a toxicological concentration of MPs, anxiety-like behavior arose, while 1 mg L−1 treatment showed a decrease in hepatic mitochondrial respiration and membrane potential (ΔΨ), both indices of suppression of mitochondrial respiratory chain [101]. Trevisan et al. (2019) also reported that NPs can cause mitochondrial energy disruption, a decline in energy efficiency, and differential mitochondrial uptake in developing zebrafish [102]. Mitochondria can also be a target for PAHs toxicity as its high lipid content facilitates PAH uptake [103]. Several studies reported that PAHs, such as benzo[a]pyrene (BaP), phenanthrene, and fluoranthene [104] or complex PAHs mixtures, can impair mitochondrial bioenergetics in embryonic or larval fish [105][106]. Gu et al. (2023) found that in sea cucumber (Apostichopus japonicus), exposure to polystyrene nanoplastics (PS-NPs) significantly inhibited the complex activities in the mitochondrial respiratory chain and affected the relative expression levels of mitochondrial apoptosis-related genes [107].

2.4. MPs/NPs’ Effects on the Redox Homeostasis of the Aquatic Organisms

A previously mentioned, MPs and NPs are able to affect the redox homeostasis. However, the effects reported in literature sometimes are conflicting, even in the same species. Probably, the differences found are related to age [108].

2.4.1. NPs

In zebrafish embryos, at 96 h post-fertilization (hpf), while exposed 50 nm (1 mgL−1) to PS-NPs, Bhagat et al. (2022) found that ROS production was increased without a significant change in the malondialdehyde (MDA) concentration. Instead, the antioxidant system was affected with decreased SOD and GR activity, increased CAT activity, and reduced glutathione content without significant change in GPX activity [109]. In 96 hpf zebrafish embryos exposed for 24 h to 50 nm and 1 µm PS-NPs (which are considered NPs by the authors) (10 mg/L), an increase in ROS content with both sizes was found [110].
In adult zebrafish exposed to 70 nm PS-NPs (0.5 ppm and 1.5 ppm) for 7 days, a significant increase in muscle (but not liver) ROS content was found only at the highest concentration [111]. According to these results, in adult zebrafish (F0) exposed to 42 nm PS-NPs added to the diet (10% of the food by mass) for one week, no difference in either genders was found in the liver GR, GPX, and CAT antioxidant enzymes [112]. Instead, a significant decrease was found in the GR activity of male muscle. In F1 96 hpf larvae from exposed maternal and co-parental groups displayed significantly lower total thiol levels (reduced forms of protein and non-protein thiols). GR activity was significantly reduced in these larvae, while the activities of GPX and CAT were not changed [112]. In another study, Aliakbarzadeh et al. (2023) exposed adult zebrafish for 45 days to PS-NPs of different sizes, 20–80 nm, at different concentrations (0.1 µgL−1, 1 µgL−1, 10 µgL−1, and 100 µgL−1). They found an NP-concentration-dependent decrease in CAT activity and total GSH content [113]. In another study conducted on the adult zebrafish exposed to PS-NPs (103–113 nm) for 14 days at 10 µgL−1 and 100 µgL−1, a reduction in gills’ antioxidant enzymes activity (SOD and total antioxidant capacity (TAC)) was found [114]. Although it may appear that in adult zebrafish, exposure to the different concentrations of PS-NPs reduces antioxidant capacity by promoting oxidative stress, in another study, it was shown that after 3 weeks of exposure to 70 nm PS-NPs at different concentrations (20 µg/L, 200 µg/L, and 2 mg/L), there is an increase in the activity of CAT and SOD only at the higher concentration (2 mg/L) [115].

2.4.2. MPs

In 96 hpf zebrafish embryos exposed to PS-MPs 1–5 µm (2 mgL−1), the redox homeostasis results unaffected [116]. The lack of effects of PS-MPs on the redox homeostasis of zebrafish embryos was also highlighted in another work, in which, following seven days of exposure of 6dpf (days post fertilization) zebrafish to 5 μm PS-MPs (50 ngmL−1), no changes were found in the concentration of MDA and the activity of SOD, CAT and GPX [117]. Similar results were found in 30dpf zebrafish exposed to 5 μm PS-MPs (500 μgL−1) for 25 days [118]. However, while in 96hpf zebrafish, the activity of antioxidant enzymes was not affected by MPs exposition, in 30 dpf zebrafish, CAT activity was decreased [118], reinforcing the hypothesis that the MP-induced impairment of redox homeostasis is age-related. On the other hand, Guimarães et al. (2021) found that in juvenile zebrafish exposed to two different concentrations of PS-MPs (4 × 104 and 4 × 106 particles/m3) for five days, an increase in thiobarbituric acid reactive species (TBARS) following both treatments. However, they found no increase in the hydrogen peroxide concentration but verified an increase in the SOD and CAT activity enzyme and in the total glutathione concentration, without difference in the reduced glutathione concentration [108].
CAT and SOD were also increased in the gut tissue of adult zebrafish exposed to PS-MPs (5 µm) for 21 days at two different concentrations (50 µgL−1 and 500 µgL−1) [119]. Another work conducted on adult zebrafish exposed to PS-MPs 9–10 µm (10 mg/L) for 4 or 8 days reports an interesting view of the effect exerted by MPs on the liver [120]. They found that after four days of exposition to MPs, the activity of the SOD, CAT, and GPX enzymes was increased, while MDA concentration was unaffected. After eight days of exposition, SOD activity returned to the control level, while CAT and GPX activity remained higher, and MDA content decreased compared with the control. Accordingly, Lu et al. (2016) found that after three weeks of exposure of adult zebrafish to PS-MPs (5 µm) at different concentrations (20 μgL−1, 200 μgL−1, and 2 mgL−1), SOD activity increased at all concentrations, while CAT activity increased with 200 μgL−1 and 2 mgL−1 treatment [115].

This entry is adapted from the peer-reviewed paper 10.3390/environments10090161

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