Consequences of nano and microplastic exposure in rodent models: the known and unknown

The ubiquitous nature of micro- (MP) and nanoplastics (NP) is a growing environmental concern. However, their potential impact on human health remains unknown. Research increasingly focused on using rodent models to understand the effects of exposure to individual plastic polymers. In vivo data showed critical exposure effects depending on particle size, polymer, shape, charge, concentration, and exposure routes. Those effects included local inflammation, oxidative stress, and metabolic disruption, leading to gastrointestinal toxicity, hepatotoxicity, reproduction disorders, and neurotoxic effects. This review distillates the current knowledge regarding rodent models exposed to MP and NP with different experimental designs assessing biodistribution, bioaccumulation, and biological responses. Rodents exposed to MP and NP showed particle accumulation in several tissues. Critical responses included local inflammation and oxidative stress, leading to microbiota dysbiosis, metabolic, hepatic, and reproductive disorders, and diseases exacerbation. Most studies used MP and NP commercially provided and doses higher than found in environmental exposure. Hence, standardized sampling techniques and improved characterization of environmental MP and NP are needed and may help in toxicity assessments of relevant particle mixtures, filling knowledge gaps in the literature.


Introduction
Plastic debris is a growing environmental concern. In 2019, 368 million tons of plastic were produced globally [1]. Furthermore, pandemic-related single-use plastics (i.e., surgical masks) have worsened the scenario [2]. Despite recycling initiatives and legislation to ban singleuse plastics, different plastic particles have been found in oceans, fresh water and agricultural systems, urban environments, the atmosphere, and remote areas such as the Mount Everest [3][4][5]. Small plastic particles are defined as microplastics (MP) (less than 5 mm diameter) and nanoplastics (NP) (less than 100 nm) [6,7] and can vary in size, shape, type of polymer, and concentration [1,3,8,9]. Regarding the sources, these are either deliberately manufactured (primary MP/NP) or derived from larger plastics during environmental exposure such as UV irradiation, mechanical abrasion, or microbial degradation (secondary MP/NP) [8].
Plastic particles are far-reaching and a multifaceted problem. The focus is not only on food [10,11] or aquatic systems [4,7,8] as primary sources of plastic exposure but also on its epidemiological consequences [9,[12][13][14]. Small volume but large surface area facilitates chemical reactions with body fluids and tissues in direct contact with particle surfaces. These particles are of particular concern due to their persistence, bioaccumulation in the food chain and in wildlife destined for human consumption, potential toxicity, and ability to act as vectors for pathogens and co-pollutants [9,12]. Marine organisms have also presented toxic effects of MP and NP exposure, depending on the type of organism, ultimately affecting bioaccumulation, metabolic changes, inflammation, reproduction effects, behavior, and ecosystem interactions [8,15]. In addition, fish exposed to NP by environmentally relevant exposure route (contaminated prey ingestion) showed NP accumulated in different fish tissues and affected innate immune gene signatures. This exposure may compromise their ability to survive in nature [16]. Humans are exposed either directly to MP and NP in drinking water, sea salt, and the atmosphere or indirectly through the food chain [8][9][10][11]. Debris from plastic prosthetic implants is also a source of exposure to MP and NP in humans [9]. Moreover, the accumulation of particles in all trophic levels may expose humans to more particles in food sources [10,13]. In a recent systematic review about MP content in American food sources, a caloric intakebased calculation was used to estimate human ingestion of a large number of particles (> 50,000) per year, significantly rising if drinking bottled water was included [17]. Such studies are necessary to raise public awareness about the constant uptake of plastic into the human body. It remains a matter of debate, however, which types of particles or their size or cargo as well as location may be critical in driving specific health-related conditions and diseases.
Continuous sources of less-concentrated MP (food containers and drinking water) are also a concern. Regulators (EFSA/WHO) state that MP and NP exposure in humans present few adverse effects, although this statement may be due to little evidence rather than a lack of effects. Preliminary signs of harm are still arising. The precautionary principle recommends and supports initiatives to develop better analytical methods before concluding that MP and NP exposure is entirely safe after all [18]. Current estimations of plastic particle exposure in humans are limited due to the lack of an established method to provide non-destructive evidence of MP and NP presence in tissue [10]. Ultra-thin sections of tissue, often used in medical research, cannot clarify the possible involvement of plastic in disease processes, as plastic is technically challenging to identify due to its small size and chemical inertness. Assessment of MP and NP exposure in rodent models offers a valuable tool to assess health risk of plastic exposure to animals and parallel it to humans. In addition, many established rodent models of human diseases offer the possibility to assess the sensitivity of specific pathologies to MP and NP exposure. We review recent findings from MP exposure within in vivo rodents model systems intending to give an outlook on them beyond the highlighted gastrointestinal and respiratory tract possible effects and fill knowledge gaps within other systems as well.

Searching methods
In this scoping review, we used different combinations of keywords in the Google Scholar database between 2001 and 2021: "microplastics"; "nanoplastics"; "exposure"; "oral administration"; "inhalation"; "rodents"; "mice"; "rats"; "accumulation"; "toxicity" and "toxic effects". Inclusion criteria were original studies published in peer-reviewed journals and performed by exposing rodents (mice and rats) to MP and NP, assessing the accumulation of particles in tissues and/or toxic effects. With that, 31 original studies were included and described in Table 1. The remaining manuscripts were included as complementary information.

Plastics utilized in rodent models
Plastics are synthetic polymers derived from fossil fuels or biomass. The most common polymers produced globally include polyethylene terephthalate (PET), polyethylene (PE), polyvinyl chloride (PVC), polypropylene (PP), polystyrene (PS), and polyurethane (PUR) [19]. Heterogeneous plastic mixtures contaminate environmental sources such as water [20,21], in which environmental fragmentation and degradation may hinder their classification, generating products with different shapes, sizes, chemical compositions, and densities [14]. However, most rodent studies used one plastic entity (Table 1) and not with heterogeneous mixtures as found in the environment.
Commercially available particles are uniform spheres with pristine or functionalized surfaces. Despite the characterization of exposure effects of a particular polymer, commercial specifications do not reflect environmental exposure accurately [14]. To this end, Estrela and colleagues assessed acute exposure to the combination of zinc oxide nanoparticles and PS NP in mice [22]. Although pathophysiological changes were observed from exposure to PS NP (Table 1), no additive or synergistic effects were observed when administered in combination. Moreover, Liang and colleagues found that MP and NP mixtures with different sizes facilitate biodistribution in mice's tissues [23].
Secondary MP and NP exhibit diverse shapes and surfaces from environmental weathering that may influence biodistribution. For example, an assessment of tritiated polyethylene glycol (PEG)ylated PS in a tumor model nude mouse highlighted the accumulation of rod/worm-like particles in the liver and spleen compared with retention of small spherical particles in tumor masses [24]. However, further work is needed to determine the effects of polydisperse environmental secondary particles. In addition, the development of improved sampling methods to accurately characterize 'natural' particles is necessary [20,21]. According to our literature review, label-free determination of plastic in cells and human-relevant systems has not yet been successful, although innovative microscopic or spectroscopic methods (e.g., UV light spectrum, infrared light spectrum, and Raman spectrum) are still emerging [25]. Radio-labeled plastic particles are used to include quantitative whole-body radiography in marine organisms and determine the mass balance in mice [24,26,27]. Fluorescently-labeled MP and NP facilitate direct quantification of bioaccumulation in tissues. Also, many commercial particles are produced with internalized fluorescence, avoiding dyespecific interactions on the particle surface. Nonetheless, possible effects of label leaching over time must be considered [28,29].
Quantifying particle deposition within tissues helps determine whether responses are due to direct interactions with particles or indirect secondary effects [28,29]. Monitoring labeled polymers non-invasively offer the potential for real-time measurements. For instance, Amereh and colleagues observed the accumulation of a mixture between 25 and 50 nm polystyrene particles in testes of Wistar rats using in vivo imaging system (IVIS) [30]. Another study using IVIS showed accumulation over time in the intestines of mice exposed to MP and NP [23]. However, longitudinal monitoring of fluorescent probes is hampered in deep tissues by signal penetration and tissue autofluorescence. Also, due to the low resolution, positive fluorescent signals are likely to be aggregates rather than being dispersed particles. Those difficulties may justify the observation of particle fluorescence only in peripheral tissues.
Plastic contaminants should not be viewed as isolated particles as several organic and non-organic molecules can adhere to them. Proteins can, for example, form a protein corona around particles [31,32]. However, it is unclear whether these are human-relevant proteins and their effect. Other toxic molecules can also bind to plastic (some of them already during the manufacture of plastic products) and are slowly released later into the environment or the body [33]. Moreover, plastic binds to lipids or changes their composition in cell membranes, which may occur in freshwater algae [34]. However, we did not find any information on such phenomena in rodents or human-relevant systems.
Due to synthetic production and environmental degradation, plastics are in close contact with several types of additives and pollutants, such as phthalates, bisphenol analogs, surfactants, and pigments, all associated with potential toxic effects [14]. For example, Deng and colleagues demonstrated phthalate ester accumulation in the gut, liver, and testes following exposure to PE MP by oral gavage [35]. Moreover, several chemicals can act as endocrine disruptors, i.e., affecting hormones pathways or acting as pseudo-hormones themselves [36,37].
In summary, improved sampling methods to determine the most common environmental particle properties will help to streamline the systematic characterization of the effects of individual polymers of different shapes, sizes, and associated coronas. In addition, the experimental utilization of heterogeneous mixtures of particle combinations and environmental plastic samples may contribute to a better understanding of the potential additive effects and effects of chemicals that come as cargo with MP and NP exposure.

Dosage
The environmental relevant dose of MP and NP exposure is heavily debated. Many studies use MP and NP concentrations far greater than current human exposure estimates (Table 1). Estimations are that human consumption of up to 0.06 mg/kg/day of particles occurs via drinking water [30]. Administration of a high single dose of particles followed by substantial recovery or constant exposure of concentrated particles is unlikely to reflect real-world scenarios. To this end, Stock and colleagues used a dosing regimen of PS MP at less than 34 mg/ kg body weight thrice weekly for four weeks [38]. They found minimal particle uptake into intestinal tissue and no toxic effects.
Conversely, high concentrations reflect the combination of multiple exposure routes in nature [39] and emulate increases in microplastic pollution in the future. Current limitations in methods to detect MP and NP accurately hinder estimations of environmental concentrations unreliable [40]. Therefore, determining the threshold at which MP and NP exposure is associated with adverse events remains critical.

Polymer exposure routes
Oral ingestion of plastic and absorption via the gastrointestinal tract has so far been the focus of MP/NP research [38]. However, reports in which plastic particles sized up to 20 µm are ingested [41] do not seem comprehensible according to the assessment of the German Federal Institute for Risk Assessment (BfR) [42]. Although microparticles up to 150 µm can translocate across mammals' intestinal barriers [43], the absorption rate is below 0.3%. From the rate, mostly particles sized up to 10 µm should be able to penetrate all organs, including the brain, with unexplored consequences [44]. Low absorption of MP and NP through intestinal epithelium could be related to particles properties and efficiency of the mucus barrier to interact and maintain MP and NP in the intestinal lumen. By being maintained, MP and NP can be excreted in the feces or deposited, which may cause local irritation or release of toxic additives [44]. Also, MP and NP can be internalized by intestinal epithelium and be re-released into the intestinal lumen due to a turnover of approximately 3 days, thus not reaching the bloodstream [45]. Currently, some studies assume that toxic effects are expected in the digestive tract and liver due to continuous plastic accumulation (Table 1) [46,47]. A murine model fed with PE particles showed increased inflammation in small intestines followed by changes in microbiota and increased systemic pro-inflammatory markers [48].
Another route for human exposure to MP and NP is drinking water, as plastic particles were detected in tap and bottled water [17]. Some studies used this administration route to expose rodents models to MP and NP (Table 1). However, water consumption was not assessed for particle intake calculations [49][50][51]. Additionally, this route is not appropriate for assessing buoyant polymers such as PP and PE and may be inefficient considering particle sedimentation over time for MP and NP suspensions. Another limitation of the oral uptake route (drinking water, diet, and oral gavage) might be bioavailability, which was estimated to range from 0.2 to 1.7% with different types of NP in vivo [52].
Plastic is not only absorbed by food through the digestive tract [53]. It can also be inhaled through fine air dust (e.g., abrasion from car tires or clothing [54,55] and release chemical additives [56] once within the body [57]). Occupational diseases associated with textiles have been extensively reviewed [54]. Fragments and fibers are the most common forms of atmospheric MP and NP. However, estimations of human exposure levels are limited by the lack of sensitivity of current methods to detect small particles [5,58].
Clearance of inhaled particles can be through mucociliary transport resulting in negligible deposition in airways or phagocytosis by alveolar macrophages or lymphatic transport [54]. MP and NP may avoid these mechanisms, accumulating in the lungs and entering systemic circulation [27,58,59]. Inhaled nanoparticles can also reach the central nervous system (CNS) through the olfactory bulb [60]. A recent 14-day repeat inhalation study in rats highlighted lung inflammation and decreased inspiratory rate following exposure to 100 nm PS particles [58]. Also, a single intratracheal dose during gestation resulted in maternal-to-fetal translocation of PS NP [59].
Topical exposure to MP and NP from microbeads in personal hygiene products and contaminated water may directly affect the skin. Epidermal cells exposed to MP and NP in vitro exhibited oxidative stress [61]. However, uptake across the outermost skin layer, the stratum corneum, is considered restricted to nanoparticles smaller than 100 nm [43]. Minimal uptake was observed following ex vivo administration of 20-200 µm fluorescent particles to pig ears both with and without compromised barrier function [62]. Particle weathering and aging may enhance topical uptake, as observed in mice with quantum dot nanoparticles [63]. To our knowledge, topical plastic exposure has not been extensively characterized in rodent models.
Various exposure routes have been utilized in animal models. Oral and inhalation routes are considered the main exposure routes in humans. The influence of a particular administration route on particle characteristics (e.g., accompanying corona or ability to release toxic chemicals) is not well understood.

In vivo effects of polymer exposure
Despite being considered chemically inert compared to plastic monomers, toxicity following MP and NP exposure was described (Fig. 1, Table 1). MP and NP toxicity may result from their persistent physical presence in tissues. Size-dependent effects have been demonstrated in vitro with PS spheres [61,64]. Small and positively charged particles may have greater bioavailability in mammals [65]. Particle accumulation has been demonstrated in organs such as the liver, kidneys, brain, spleen, and reproductive organs (Fig. 1, Table 1), although it was independent of the functionalized surface coating in high concentration [52].
Disruption, penetration, absorption, and endocytosis mechanisms, which may be toxic [66], are currently being discussed as possible ways plastic particles can enter and interact with cells and tissues [67,68]. Possible toxic consequences may not only be due to MP and NP exposure, as most commercially available particles used in studies in vivo are provided in aqueous suspensions with dispersant and conservant solvents. Walczak and collaborators centrifuged the particles for conservant and surfactant removal before usage, controlling possible effects found after exposure [52]. Thus, evaluating additional compounds as control groups and not only test vehicle solutions is essential.

Direct effects and underlying mechanisms
In mice exposed to fluorescently labeled particles, localized inflammation at the site of particle accumulation  19:28 has been confirmed in the liver [41,69,70] and testes [71]. However, fluorescent dye leaching from MP and NP could also contribute to the exposure effects observed. Interestingly, few studies evaluated the fluorescent dye leaching of particles under conditions such as simulated gastric and intestinal fluids, and fluorescence leaching was negligible [23,52]. Moreover, fluorescent MP and NP are mainly used only for bioaccumulation and biodistribution assessments into tissues, and non-fluorescent for toxicity evaluations [23,41]. Exposure to non-fluorescent particles resulted in increased inflammation in primary absorption sites consistent with the exposure route, such as the gut [48] and lungs [58]. One proposed central mechanism for MP and NP toxicity is the induction of oxidative stress, which has been extensively observed in vitro [72]. However, another study found the opposite effect, a reduction of plasticinduced oxidative stress in cells in vitro [73]. In addition, some cell types can actively excrete plastic particles [64], possibly influencing the response to oxidative stress [74]. Mice exposed to drinking water with high concentrations of MP showed impaired antioxidant defenses, such as decreased superoxide dismutase (SOD) and glutathione (GSH) expression and increased malondialdehyde (MDA) formation (a product from lipid peroxidation). In addition, increased activity of the Nrf2/Keap1 pathway was observed, suggesting plastic-induced oxidative stress and its relation with inflammation in the tissue microenvironment [69].
Regarding additives and pollutants leached from plastic particles, mice exposed to MP and NP (PS and PE) by drinking water with organic flame retardants presented more pronounced oxidative stress in the liver [70]. Testes of mice exposed to oral gavage with PE coated with phthalate esters also showed oxidative stress responses [35]. However, these effects may be due to additives released in the solution and not to MP and NP exposure, as no information was provided regarding solutions stability over time or whether they were used as fresh preparations [35,70]. Mice exposed to a single dose of MP and NP mixtures with different sizes by oral gavage showed increased ROS generation, intestinal epithelium apoptosis, and intestinal permeability, and pre-treatment with antioxidants reversed the effects [23]. Current studies do not indicate genotoxicity or mutagenicity of everyday plastics, as shown for PS [75]. In contrast, an in vitro study in human fibroblasts [76] and an investigation into the damage to cell-free DNA [77] indicated corresponding genotoxicity. However, other types of plastic and rodents models have hardly been investigated to confirm effects on a broader species scale.
Changes to the intestinal microbiota contribute to metabolic disorders, including obesity and diseases such as colorectal carcinoma [82,83]. Li and colleagues observed increased microbial load and diversity in fecal samples of mice fed with PE particles (600 µg/day for 35 days) [48]. Gut dysbiosis coincided with increased hepatic bile acid levels and altered serum bile-and amino acid-related metabolites in mice exposed to high concentrations of 5 µm PS MP (100 and 1000 μg/L) in drinking water for six weeks [49].

Hepatotoxicity
In response to oral exposure to MP and NP, multiple groups showed altered gut microbiome and disruption of serum and hepatic markers of amino acid synthesis and metabolism, energy, and lipid metabolism [49][50][51]79], followed by liver inflammation [41,69]. Hepatocellular edema and inflammatory cell infiltration were observed with increased hepatic IL-1β and TNF-α mRNA following exposure to 5 µm PS particles (20 mg/kg/day body weight) in drinking water for 30 days [69]. The extent of hepatotoxic insult was not sufficient to alter serum markers of liver function (alanine transaminase [ALT] and aspartate aminotransferase [AST]) after the exposure period. However, mice exposed to 250 nm PUR particles by oral gavage for 10 days showed increased serum ALT, alkaline phosphatase (ALP), IL-6, and TNF-α levels, followed by liver vascular congestion and hepatocytes vacuolization [84]. Accumulation quantification of fluorescent particles was hindered by extensive tissue autofluorescence, hampering to conclude whether the effects were associated with the presence of hepatic particles.
Stock and colleagues treated heme oxygenase-1 (HO-1) triple transgenic (HOTT) reporter mice with a mixture of 1, 4, and 10 µm PS particles by oral gavage [38]. These animals expressed a LacZ reporter sensitive to oxidative stress and inflammation. However, the study found no positive responses or pathological changes to the liver or other organs, possibly due to the low concentrations of particles (1.25-34.0 mg/kg body weight for particles mixture every 3 days for 28 days).
The liver is the primary site for lipid metabolism and is sensitive to pathologies such as nonalcoholic fatty liver disease (NAFLD) that manifest as an accumulation of fatty vesicles combined with elevated circulatory cholesterol and triglycerides [83]. Lipid disruption in response to MP/NP exposure in rodents has been observed by multiple groups [50,51,82,85]. Luo and colleagues observed hepatic ballooning (characteristic of apoptosis), increased hepatic triglycerides, total cholesterol, and decreased PPARα and PPARγ mRNA in maternal mice after exposure to 5 µm PS MP (100 and 1000 µg/L) by drinking water during gestation and lactation [50]. Disrupted PPAR signaling and decreased hepatic triglycerides and total cholesterol were also observed in F1 offspring. The lipid-sensitive nuclear receptor PPARα regulates fatty acid catabolism and clearance and is thought to have anti-inflammatory effects (NF-κB suppression) [86]. Therefore, the extent of hepatic PPARα downregulation is predictive of NAFLD severity.
PPARγ is also downregulated during hepatic stellate cell activation, resulting in fibrosis [86]. At lower concentrations of 5 µm PS (500 µg/L), hepatic fatty vacuoles were observed in male C57BL/6 wild-type mice exposed to MP by drinking water for 28 days, without changes to hepatic triglyceride or PPARγ at the protein level [85]. This result indicates potential strain and/or gender differences, although a lack of water intake assessment may have resulted in different particle exposure between individuals. However, Lu and colleagues observed decreased liver weights and hepatic and circulatory levels of total cholesterol and triglycerides with downregulation of hepatic triglyceride synthesis in male mice exposed to 0.5 and 50 µm PS MP (100 and 1000 μg/L) by drinking water for 35 days. At the mRNA level, increased PPARα and decreased PPARγ expression were identified [79].
Changes in lipid metabolism are thought to be dependent on particle size. F1 offspring from dams exposed to 0.5 and 5 µm PS particles (100 and 1000 µg/L) in drinking water during gestation exhibited decreased hepatic total cholesterol and triglycerides in a particle dose-and sizedependent manner [51]. In addition, decreased PPARα hepatic mRNA expression was observed in groups exposed to 5 µm MP alone. Whether these effects are due to altered maternal metabolism, making offspring more susceptible to disease, or particles transferred to the fetus directly affecting the next generation remains unclear.

Reproductive dysfunction
MP and NP have been shown to accumulate in reproductive tissues [23,30] and cross the placental barrier [59]. Accumulation of MP and NP in testes of rodents corresponded with histological changes followed by local inflammation and DNA damage in germ cells [30,71,87]. Also, rodents exposed to MP and NP by oral gavage showed decreased serum testosterone levels, a hormone essential for spermatogenic cells development [30,71,87]. These observed effects were alleviated in male mice treated with ROS scavenging compounds because oxidative stress was induced through p38 MAPK signaling pathway activation after MP exposure [87]. This pathway is also involved in inflammation, which could explain increased levels of pro-inflammatory cytokines in testes of mice exposed to MP and NP [71,87,88]. Additionally, mice exposed to MP by drinking water demonstrated increased NF-κB followed by decreased Nrf2 and HO-1 in testes, suggesting this increased pro-inflammatory profile may be due to reduced Nrf2/HO-1-mediated NF-κB inhibition pathways [88]. Plastic exposure of mice dams caused far-reaching effects on milk ingress [50] and generally metabolic syndromes [51] in first and second-generation offspring of the first and second generation, regardless of sex [89]. In ovaries, exposure to MP by drinking water for 90 days reduced the number and volume of growing follicles and anti-Müllerian hormone levels and induced oxidative stress in rats [90,91]. In addition, oxidative stress triggered cell death mechanisms, inflammation [90], and fibrosis through Wnt/β-catenin pathway activation in ovaries [91]. Changes in the uterus due to plastic exposure were also observed [92], with altered number and gender ratio of offspring of parents exposed to PE MP by oral gavage during pregnancy. However, tendencies were not dose-dependent [93].
Exposure to PE MP in dams by oral gavage during pregnancy and lactation altered the development and number of T cells in spleens in offspring of both sexes. Also, the maturation of dendritic cells was inhibited in males and enhanced in female pups [93]. Furthermore, in an allogeneic mating murine model, pregnant mice exposed to PS MP by IP administration showed increased resorption rates of embryos, decreased number and diameter of uterine arterioles, and disturbances of maternal-fetal immune microenvironment, which compromises embryos development [94].
Metabolic disorders were also observed in offspring of dams exposed to PS MP by drinking water during pregnancy [50,51] and lactation [50]. To evaluate the long-term effects of MP and NP exposure, Luo and colleagues analyzed physiological, pathological, and metabolism indicators of adult F1 offspring (40-weeks old) of dams exposed to PS MP during pregnancy and lactation. Adult female F1 offspring showed increased lipid accumulation in the liver [50]. Furthermore, pregnant mice exposed to MP and NP by IV administration showed decreased embryo body weight, although not affecting the number of embryos [95]. In addition, mice dams exposed to 60 nm NP showed decreased placental diameter and extravasation in fetuses and placenta [95].

Neurotoxicity
Nanoplastics can cross the blood-brain barrier in a sizedependent manner [96]. Bioaccumulation, altered lipid peroxidation, and disrupted activity of neurotransmitters have been reported in the brains of marine organisms and fish [96,97]. However, plastic-mediated neurotoxicity in rodents has been poorly investigated so far [97]. While no significant differences in cognitive function were observed in rats exposed to PS NP for five weeks by drinking water, the authors noted that the small sample size (n = 6) and limited testing unlikely reflected subtle, transient effects [98].
Estrela and colleagues observed impaired object recognition in response to PS NP exposure, coinciding with redox changes, reduced acetylcholinesterase (AChE) activity, and accumulation of particles in the brain [22]. Nonetheless, administration of particles systemically (IP) does not reflect the first-pass effect and is not considered a relevant exposure route for environmental MP and NP. Furthermore, altered neurotransmitter activity following MP and NP accumulation was observed in organs besides the brain, such as the liver [41], highlighting the potential for particles to damage CNS function in multiple tissues. In addition, indirect effects of particle exposure, such as pro-inflammatory mediators from other accumulation sites, may also result in neurotoxicity [99].

Other effects
The potential effects of MP and NP exposure in other tissues are still poorly investigated in rodent models. For example, rats exposed daily to MP for 90 days by drinking water showed myocardium alterations, such as vascular congestion, areas with thinner muscle fibers and ruptures, and increased serum heart damage markers (CK-MB and Troponin I) [100]. Also, increased apoptosis and oxidative stress in the heart were observed, which triggered activation of the Wnt/β-catenin signaling pathway related to myocardium fibrosis [100].
Another concern is the potential toxicity in endocrine tissues caused by plastics. For example, rats exposed daily to PS NP for five weeks by oral gavage showed decreased active forms of thyroid hormones (FT3 and FT4) in circulation and increased levels of TSH with high doses of NP, followed by changes in cholesterol serum markers and more liver damage. Hence, PS NP administration could interfere with lipid metabolism by disrupting the thyroid endocrine system [101]. The pathophysiology of chronic inflammatory diseases and co-morbidities of metabolic syndrome may be exacerbated in individuals exposed to excessive MP and NP levels. Administration of 5 µm PS particles by drinking water in a murine acute colitis model enhanced hepatic lipid disruption and intestinal barrier dysfunction [85]. Serum inflammatory markers were higher in mice with colitis than in control animals exposed to MP, indicating the potential for sensitization of individuals with substantial plastic loads to chronic diseases.

Future perspectives
New studies are continuously published regarding possible harmful effects in terrestrial mammalian organisms caused by plastic particles. However, most studies have a set of inherent challenges that need to be overcome. Considering plastic particles are found everywhere, the first challenge is the presence of contaminants during analysis. Contaminants were described in detecting plastic particles in controls, possibly from contact with air and plastic released from clothing and laboratory materials. In addition, the high diversity of plastic properties, such as insolubility to non-harmful solvents and buoyancy, can compromise the main experimental models to assess toxicity.
Another challenge is the availability of environmental plastics, like heterogeneous mixtures compared to commercially available plastics used in studies, which cannot be extrapolated to reality. This lack of studies on environmental plastics is mainly related to poor improvement in sampling, processing, and detection of plastics loads, which also compromises estimations of MP and NP doses found in the environment. This issue converges with another challenge: doses applied in many in vivo studies do not correspond to plastics concentrations found in the environment, and studies using environmentally relevant doses showed no effects, diverging from high doses experiments.
Many variables and conditions are applied in different studies designs; thus, considering multiple testing problems that could be related to data and performing proper adjustments for each case is needed for satisfactory conclusions and suggestions. Studies may use the precautionary principle as an argument for evaluating exposure to high doses of MP and NP before assessing the environmental dose. However, literature bias may occur for publications demonstrating effects, conflicting with the studies using low doses, as they might show different results or absence of effects. Furthermore, low incentives for studies with no effects may further compromise a critical debate regarding exposure to MP and NP.
Although plastics are compounds that can be in nature for a long time, longitudinal monitoring for plastic toxicology remains poorly explored. Experimental chronic models assessing only one terminal endpoint may not show effects, hence questioning the exposure period required to observe effects. Additionally, improvement in experimental designs for long-term and chronic studies may help comprehend immunogenic responses to prolonged plastic exposure.
Several knowledge gaps were addressed in this review: synergistic or antagonistic effects of particle mixtures on uptake, biodistribution, bioaccumulation, clearance, and biological responses; standardized method(s) of assessment of particle combinations or environmental plastics is vital for appropriate risk assessment of reliable exposure concentrations and time; lack of non-invasive or non-destructive estimation of particle load and biodistribution at an adequate resolution. These knowledge gaps may be filled by improving sampling, processing, and detection in optimal resolution, leading to better estimations and the development of experimental designs closer to the environment.

Conclusion
Understanding cytotoxic effects of plastic exposure requires more progress in several fields. First, standardized sampling techniques and improved characterization of environmental MP and NP are needed. Second, will there is a good body of evidence on acute plastic exposure, chronic exposure over longer time frames in higher organisms is understudied. Third, consensus on the effects and methodological tools on the presence of plastic in vertebrates in different types of organs are lacking to better understand potential relationships to chronic inflammation and disease. More research is needed to shed light on those aspects to better understand the consequences of plastic exposure in human health and environmental risks.