Marta Llorca
Marinella Farré- Department of Environmental Chemistry, IDAEA-CSIC, Barcelona, Spain
Humans are exposed to micro and nanoplastics (MNPLs) through inhalation, ingestion and, to a lesser extent, dermal contact. In recent years, new insights indicate the potential of MNPLs to cause damages to human health. Particle toxicity can include oxidative stress, inflammatory lesions, and then increased internalization or translocation through tissues. On the other hand, plastic additives are used in plastic particles, once internalized, can release toxic substances. It is noteworthy that the potential effects of MNPLs encompass a wide range of polymers and chemical additives, showing various physicochemical and toxicological properties, and the size, shape and surface properties are other variables influencing their effects. In spite of the research carried out recently, MNPLs research is in its early stages, and further investigation is required. In this review article, the knowledge of human exposure routes and the recent results on the toxicological effects of MNPLs in human health are presented and discussed. Finally, the current limitations and the main gaps in the body of knowledge are summarised.
Introduction
Fossil-based plastics have been one of the most impressive revolutions of the last century, contributing to improving the quality of life of humans. Nowadays, plastics are used in all of the industrial sectors because there is a wide variety of available versatile materials. They are inexpensive, strong, durable, lightweight, resistant to degradation, and present thermal and electrical insulation properties (Andrady and Neal, 2009). For that reason, the global production reached 368 million tonnes in 2019 (Plastics the Facts, 2019), and this trend is expected to continue in the coming years. However, as a consequence of incidental events and mismanaged wastes, plastic residues of different sizes can reach the environment and as a conseuence they are currently ubiquitous contaminants. While the problems caused by microplastics (MPLs) in the environment (Eerkes-Medrano et al., 2015; Horton et al., 2017), and in particular in marine ecosystems (Cole et al., 2011; Ivar Do Sul and Costa, 2014; Llorca et al., 2020a), is a well recognised problem, more recently there is increasing concern regarding the potential for human exposure to MPLs (Wright and Kelly, 2017). Due to the use of items made of plastic and the indirect impact of plastic pollution, humans are constantly exposed to micro- and nanoplastics (MNPLs) through the inhalation of fibres from clothing and fabrics, tyre dust from motor vehicles, ingestion of water with residues (from filters, pipes, and bottles), ingestion of contaminated foods, among many other routes of exposure (Prata et al., 2020).
MPLs have been defined as plastic pieces with a size below 5 mm (Besseling et al., 2013), including plastic particles at the nano range, known as nanoplastics (NPLs) (Bouwmeester et al., 2015). MPLs can have their origin in cosmetics and cleansing products, cloth fibres and vehicle tyre erosion (classified as primary MPLs), and the environmental fragmentation and erosion of plastic pieces and debris (classified as secondary MPLs) (Andrady, 2011). It is important to highlight that the toxicological impact of plastic pollution is partially influenced by the size and shape of the particles (Hwang et al., 2020), as is the case with other nanomaterials (Bouwmeester et al., 2015), as they have much smaller particles and greater ability to be internalised and translocated through tissues in living organisms. As a consequence, their potential toxicity also increases. Nothwithstanding, the smaller the particles, the greater their active surface area, hence they facilitate the leaching of plastic additives, which are chemical ingredients that are added to polymers to enhance their performance and characteristics, but they are not chemically bound to the polymeric chain. Therefore, they can easily leach from plastic particles being another route of exposure to these chemicals in addition to others. Moreover, some plastic additives have already been shown to have toxicological properties that are harmful to the environment and human health (Hermabessiere et al., 2017). On the other hand, the nature of the most used polymers facilitates that once in the environment MNPLs can adsorb and facilitate the transport of other environmental contaminants (Llorca et al., 2018; Llorca et al., 2020a) and pathogens to living organisms, which is known as the Trojan horse effect. This ability to adsorb and desorb other contaminants and pathogens is not only dependent on the polymeric composition of the particle, but is also size dependent because, again, the smaller the particles, the greater their active surface area.
However, to the best of our knowledge, to date the consequences of MNPL exposure on human health are not well understood. In addition, the extensive number of polymers and plastic additives that are involved, with variables such as the size and shape of the particles, and the current analytical limitations, should also be considered.
Under this frame this paper has been organised into the following sections: 1) main human routes of exposure; 2) pathways of particle toxicity; 3) toxicity of plastic additives on human health; 4) insights about Trojan horse effects on human health; and 5) future reaserch trends.
Main Routes of Human Exposure
Recent studies indicate that MNPLs are present in most classes of consumer products. In foodstuffs, MNPLs can be present in animals which are contaminated through their environment or food chains, as happens with seafood (Santillo et al., 2017) for example, and also they can be contaminated during their production processes or packaging (Lau and Wong, 2000; Mason et al., 2018; Du et al., 2020). In the case of drinking water, MPL contamination can come from pipes, filters, or bottles (Oßmann, 2021). In addition to a portion of MNPLs that we can ingest incidentally, such as the MPLs that are present in some toothpaste formulas. Also, dust from plastics, car tyres, paints, and textile fibres are sources of MNPLs in the Earth’s atmosphere and can be inhaled or can undergo dermal interaction with humans. Moreover, MPLs are used in the formula of certain personal care products (PCPs) (Laborda et al., 2021), implying dermal interaction. Therefore, the main routes of human exposure to MNPLs are ingestion, inhalation, and dermal contact (Figure 1). However, the total contribution of MNPLs to exogenous particles exposure is still unknown, and due to their unique composition and properties, the human biological responses will differ from those of other exogenous materials.
FIGURE 1. Summary of the human routes of exposure to MNPLs and their potential effects on health.
Ingestion is considered as the major route of human exposure to MNPLs (Galloway, 2015), and the tissues of the human gastrointestinal tract (GIT) are considered among the most exposed. In particular, the Peyer’s patches of the ileum in the small intestine have potentially been defined as the main sites of uptake and translocation of anthropogenic particles. The Peyer’s patches are part of the gut-associated lymphoid tissues (Powell et al., 2010). The M cells of these patches can transport particles (0.1–10 μm) from the intestine to the mucosal lymphoid tissues, responsible for initiating immune responses to specific antigens that are encountered along all mucosal surfaces. The subepithelial area of the Peyer’s patches can store non-degradable particles, and consequently they disrupt the immunity processes (Campanale et al., 2020). For bigger particles, up to 130 μm diameter, persorption performed by GIT epithelial cells is another route of uptake but not of adsorption and is the cellular passage of large particles through the epithelial layer of the GIT. In addition to particle size, other factors that can influence the uptake and translocation of MNPLs are their chemical composition, shape, and hydrophobicity. Hydrophobic surfaces can be better transported through the mucus layer. Moreover, these characteristics also influence the adsorption of proteins and other biomolecules to the particle surface, known as corona formation (Lunov et al., 2011). Therefore, the surface charge of MNPLs can also influence the extent and pathway of translocation to other organs. Food and drinking water are key factors in the assessment of human MNPL exposure through ingestion, but most of the studies till now have been focused on fish and seafood, rather than on other foodstuffs.
MPLs in molluscs, such as mussels (Van Cauwenberghe and Janssen, 2014; Li et al., 2015; Smith et al., 2018), have been widely studied because they are living in intertidal areas, where the abundance of MPLs is expected to be higher due to the density of the polymers. For that reason, mussels are particularly exposed (Coppola et al., 2020). Moreover, the edible part of molluscs includes the stomach, intestine and digestive glandular, and they are consumed worldwide, thus potentially representing a source of MPLs in the human diet. As shown in Table 1, the concentrations of MPLs in molluscs present a high variability, for example, ranging from 0.05 to 3 MPLs items per g wet weight (ww) of edible tissue in mussels. This variability among molluscs is highly influenced by the habitat between different species and, in particular, the degree of contamination of the different sampled areas, in addition of the diversity of analytical methods used, including the diversity of the sample preparation approaches. Karlsson et al. (Karlsson et al., 2017) studied the accumulation of MPLs in mussels and the surrounding water and sediments, and their findings indicated that MPLs in mussels were approximately a thousand-fold higher, per unit weight or volume than in the surrounding sediment and water, showing the potential of bioaccumulation of filter-feeding organisms.
TABLE 1. Summary of studies assessing MNPLs in foodstuffs.
As can be seen in Table 1, the prevalent morphotypes were microfibres. Regarding polymers, the most detected were polyethylene (PE), polypropylene (PP), polyethylene terephthalate (PET), and the polyamide group (PA).
On the other hand, despite the number of studies on MPLs in the GIT of fishes (Schirinzi et al., 2020; Abidli et al., 2021; Hadibarata et al., 2021; Ghosh et al., 2021; Li et al., 2021; Zhang et al., 2021), only a reduced number of studies considered the edible parts. Daniel et al. (2020) (Daniel et al., 2020) recently carried out a study in which they compared the MPLs in edible and non-edible parts of pelagic fish for human consumption, showing that MPLs in the GIT of fishes was much higher than in edible parts, such as muscle. In another study, Barbosa et al. (2020) (Barboza et al., 2020) studied the distribution of MPLs in the GIT, gills and dorsal muscle of 3 fish species, Dicentrachus labrax, Trachurus, Scomber colias, and 49% had MPLs, while 32% had MPLs in dorsal muscle, with a total mean (±SD) of 0.054 ± 0.099 MPLs items/g. The results again showed higher results in the GIT and gills than in the edible parts of the 3 studied species. 1.3 ± 2.5; 1.0 ± 1.9; 1.2 ± 1.6 MPLs per individual in the GIT, 0.8 ± 1.4; 0.7 ± 1.4; 0.7 ± 1.0 MPLs per individual in the gills, and 0.4 ± 0.7; 0.7 ± 1.3; 0.6 ± 0.8 MPLs per g, in the dorsal muscle of D. labrax, T. trachurus, and S. colias, respectively. In this regard, it should be pointed out that only the smallest particles can be internalised and translocated, while to date due to technical reasons, most of the studies are not considering the evaluation of NPLs,. Therefore, there is an important gap in the knowledge about NPL contamination in the edible parts of fishes.
Another source of MNPL contamination can be packaging. Nevertheless, there are scant studies that have considered the potential of MNPL transfer from plastic cans (holding drinks or food) and plastic films. Recently, Akhbarizedeh et al. (Akhbarizadeh et al., 2020a) assessed the abundance and composition of MPLs in canned fish. This study showed that 80% of samples had at least one MPL item, with fibres being the most abundant example. PET was present in 32% of the samples, which was the most commonly found polymer (Akhbarizadeh et al., 2020a). In another study (Karami et al., 2018), MPL contamination was studied in sardine (Sardina pilchardus) and sprat (Sprattus). In this case, 20 brands of canned sardines and sprats that had been collected in 13 countries were examined. MPLs were confirmed in only four brands, within which, between 1 and 3 MPL particles were found, and polypropylene (PP) and polyethylene terephthalate (PET) were the identified polymers.
In addition to seafood and fish, drinking water is a significant potential source of MNPLs in the human diet. As shown in Table 1, several studies reported the presence of MNPLs in potable tap water. In these studies, both fibres and fragments had been identified. Despite the fact that PVC is one of the most used polymers in pipes was marginally detected in tap water, while polyethylene (PE), Polystyrene (PS), PET and fibres of polyester were those that were more frequently confirmed. In most of the studies, MNPLs in tap water were present in frequencies superior to 90%. However, the concentration of the particles was very variable between the different studies. In this case, the results were highly dependent on the different analytical approaches used, ranging from optical microscopy, scanning electron microscopy (SEM), techniques based on Fourier Transform Infrared (FTIR) spectrometry, and Raman spectrometry, and their capacity to measure the range of the smallest particles. Harmonised methods are urgently needed for drinking water and other food matrices (Oßmann, 2021). Moreover, in the case of drinking water it is particularly relevant to measure the range from NPLs to a few μm, as well as the combination with other analytical techniques that are able to provide polymer concentrations per litre of water, such as liquid chromatography coupled to high-resolution mass spectrometry (LC-HRMS) (Schirinzi et al., 2019) or gas chromatography pyrolysis mass spectrometry (GC-Pyr-MS) because concentrations in terms of the mass of polymer in water can differ in the counting particles of each polymer, and other techniques are not able to provide an assessment of NPLs. Nevertheless, from the toxicological perspective, especially the smallest MNPLs (<1.5 μm) might be toxicologically relevant, in agreement with the European Food Safety Authority (EFSA) (Efsa, 2016).
To carry out the toxicological assessment, often only particle numbers and sizes are used, and only sometimes also concentrations in mass per litre are considered. Recently, the World Health Organization (WHO), in its report on MPs in drinking water (Who, 2019), revisited this issue.
As happens in tap water, high variability was found when comparing different studies on bottled water (Akhbarizadeh et al., 2020b). The most common polymer that was found in bottled water was PET, which is generally used for bottle production and also PP, which is used for the caps. Therefore, these data suggest that the contamination is at least partially coming from the packaging and/or the bottling process itself.
Another studied foodstuff matrix was sea salt, but as happens with water, the comparison of the different studies cannot be made, in particular, because the different studies were carried out using different techniques (optical microscopy, FTIR, GC-Pyr-MS), and some of these techniques, such as optical microscopy, only provide information on bigger particles.
A minor number of studies considered other food matrices such as soft drinks (Shruti et al., 2020a). Thus, very few studies estimated the intake of MPLs through the diet. Focussing on the diet in the United States, Cox et al. (2019) (Cox et al., 2019) estimated the intake of MPLs. Evaluating approximately 15% of American’s caloric intake, MNPL consumption ranges were from 39,000 to 52,000 particles/person per year depending on age and sex. Additionally, individuals who drink only bottled water may be ingesting an additional 90,000 MNPLs annually, compared to 4,000 MNPLs for those who consume only tap water. Notwithstanding, food preparation and cooking can influence the average MNPL concentrations in meals (Rist et al., 2019). Thus, the comparison between raw foods and final meals is required to assess human exposure to MPLs through the diet.
Inhalation
MNPL inhalation is considered to be one of the main routes of human exposure. Once inhaled, MNPLs reach the respiratory epithelium and, according to the lessons learned from particulate matter and nanomaterials investigations, it is expected that they may translocate via diffusion, direct cellular penetration, or active cellular uptake through endocytic and phagocytic processes. In the alveoli, phagocytosis is the main pathway for the particles with sizes of between 1 and 3 μm, while the smallest particles could be passively transported via diffusion across membrane pores (Wright and Kelly, 2017). Nonetheless, the number of studies assessing MNPLs in indoor and outdoor environments continues to be scant (Huang et al., 2020). Most of the studies that have been carried out to date were based on the use of μ-FTIR, which limits the assessment of NPLs. These approaches are, in general, composed of sample collection using passive or active samplers with quartz fibre filters (pore size: 2 μm, diameter: 46.2 mm). In most of the sample pre-treatment procedures, after collecting the total suspended particulate, the samples are digested to remove the organic materials. The most widely used digestion approach is the use of a H2O2 solution of 30% (v/v), sometimes with heating to speed up the process. Similar approaches are also well used in the assessment of atmospheric fallout but, in general, they are using passive collectors with different types of filters such as glass fibre filters (Dris et al., 2015), quartz fibre filters (Dris et al., 2016), nitrocellulose filters (Zhou et al., 2017), and PTFE filters (Allen et al., 2019).
As shown in Table 2, fibres are generally the most common MPLs that are identified in outdoor atmospheres. The analytical method that is employed can limit the range of particles that can be assessed. Nowadays, there is a continued existence of a critical gap of knowledge about the minor-sized particle fraction, which can be inhaled. On the other hand, relatively few research articles have considered indoor environments. In order to assess human exposure to indoor airborne MPLs, a breathing thermal manikin was used by Vianello et al. (Vianello et al., 2019) and those authors investigated MPLs down to 11 µm particle size. The manikin was used to investigate 3 different habited apartments, and the concentrations were found between 1.7 and 16.2 particles m−3, but only 4% were MPLs, with polyester being the predominant polymer in 81%, followed by PE in 5%, and PA in 3%. Another relevant result of this study was that the MPLs were typically of a smaller size than the non-synthetic particles. Zhang et al. (2020a) (Zhang et al., 2020a) studied MPL fallout in different indoor environments, and they established that dormitories were the rooms with a higher abundance of MPLs (average 9,900 MPLs/m2/day). In addition, airflow turbulence by the use of an air conditioner increased the resuspension of MPLs. In another study, the presence of MPLs in indoor dust was investigated in the city of Surabaya (Indonesia). The deposition was assessed in an apartment, an office, and a school (Bahrina et al., 2020). The greatest concentration was found in the office with an average of 1,186.36 particles/m2, and most of the MPLs were fibres. PET, polyester, and cellophane (CP) were the main polymers detected. Recently, in another study, thirty-two airborne indoor deposited dust samples were studied in homes in Sydney (Australia) (Soltani et al., 2021). MPL fibre deposition rate ranged from 22 to 6,169 fibres/m2/day, and most of the deposited dust were fibres (99%). The majority were natural fibres (42%), 18% were transformed natural fibres, and 39% were MPLs. PE, polyester, PA, and PS were found in higher abundance in homes with carpets. While when the carpet was absent, the polyvinyl fibres were predominant. Mean inhaled MPL weight was estimated to be 0.2 ± 0.07 mg/kg-body weight (BW)/year and 12,891 ± 4,472 fibres/year. It is noteworthy that the greatest inhalation intake rates were for the age group below 0.5-years, at 0.31 mg/kg-BW/year. As can be seen in these studies, there is a common trend indicating that fibres have predominant shapes. However, the rate of deposition and the rates of inhalation can be extremely different according to many factors, such as climate, the use of carpets or an air conditioner, the use of vacuum cleaners, types of textiles, number of habitants in the same place (Pratiwi et al., 2020), among many others, can affect the final rates.
TABLE 2. Summary of studies assessing MNPLs in airborne particulate and atmospheric deposition.
Dermal
Despite the number of personal care products such as scrubbing powder cleaners, shampoo, facial make-up, and balsams containing MPLs and microbeads in their formulations, and the fact that it has been suggested that there is a potential for NPLs to cross the dermal barrier (Revel et al., 2018), little attention has been paid to assessing dermal exposure. In one of the first studies to assess the potential cytotoxicity of MNPLs, Schirinzi et al. (2017) (Schirinzi et al., 2017) evaluated the effects on human epithelial cells of PE and PS, showing that PS presented Reactive oxygen species (ROS) generation. This is another indication of the importance of assessing the different plastic polymers, rather than considering MNPLs of different materials under the same umbrella. MPLs, and in particular NPLs due to their small size, 3D structure, and apolarity of the majority of polymers, can be internalised by cells (Xu et al., 2019). Moreover, the potential interaction between NPLs and proteins could lead to structural changes of proteins (Hollóczki and Gehrke, 2019) as well as the disruption of the lipid bilayer (Hollóczki and Gehrke, 2020).
Pathways of Micro and Nanoplastics in Human Toxicity
MNPLs have a complex nature involving a polymeric matrix with additives such as plasticisers, flame retardants, fillers, UV stabilisers, coating finishers, colourants, metals, among others. Thus, firstly, MNPL toxicity should consider the physical particle damages that can cause inflammatory lesions, originating from the potential of their surface to interact with the tissues. Secondly, MNPL toxicity should also consider the long-term toxicity produced by the plastic additives, which are not covalently bound to the polymer and, once internalised, could be easily released.
Nowadays, most of the information about MNPL toxicity impacting human health is from a limited number of in vitro studies and relevant gaps of information exist. According to the information gathered to date, the Science Advice for Policy by European Academies (SAPEA) project in 2019 considered the lack of evidence on direct adverse effects of MPLs on human health (Sapea, 2019). Furthermore, considering the number of current uncertainties, the European Commission Scientific Advice Mechanism also considered the need to research this field in order to obtain a global perspective compared to other contaminants (not available in Crossref, 2018). As shown in Table 3, some effects cannot be overlooked, and some are dependent on the plastic particle composition, size, and shape. In order to assess the potential toxicity of MNPLs, the capacity of cell internalisation is a central issue. Liu et al. (2021a) (Liu et al., 2021a), using a model cell membrane and rat basophilic leukaemia (RBL-2H3) cells, evaluated the cellular internalisation and release of PS-MNPLs. In that study, both models were exposed to 50, 500, and 5,000 nm particles. Those authors showed that 50 and 500 nm particles were absorbed on the model membrane due to hydrophobic interactions and Van der Waals forces (meaning dispersion forces). This range of particles was also internalised into living cells via both 1) passive membrane penetration because of the partition of PS-MNPLs in the water-phospholipid system and 2) in the case of 50 nm particles by active endocytosis through the clathrin and caveolin-mediated pathways and mainly by micropinocytosis. As expected endocytosis opens the plasma membrane to Ca2+ influx which could cause the mitochondria dysfunction. In contrast, bigger particles (5 μm cannot be internalised due to their size. The endocytosed 50 and 500 nm particles mainly accumulate in the lysosomes. In addition to cell internalisation, these particles were also excreted. Finally, it was concluded that masses of the 50 nm particles that were internalised and excreted were both higher than the masses of 500 nm (Liu et al., 2021a).
TABLE 3. Selection of toxicological studies using human cell-lines.
Since ingestion is considered to be the first mode of human exposure, the gastrointestinal tract is one of the first exposure tissues. For this reason, several studies have been carried out to explore the responses in the human colon adenocarcinoma Caco-2 cells as a model. Wu et al. have studied whether cytotoxicity of PS particles was size-dependent in exposing Caco-2 cells to PS-MNPLs of 5 μm and 100 nm (Wu et al., 2019). In that study, both particle sizes exhibited low toxicity on cell viability, oxidative stress, and membrane integrity. However, the mitochondrial membrane potential was disrupted by both, and on the contrary, as expected, the 5 μm particles induced higher effects than those of 100 nm (Wu et al., 2019). In this case the authors explained that might be due to the different mechanisms of induced mitochondrial depolarization. It was found that 0.1 μm PS-MPs accumulated in lysosomes, while 5 μm PS-MPs analyses did not reveal accumulation in lysosomes. Large particles at a micrometer scale could escape from the lysosomes after endocytosis and localize in the intracytoplasmic vacuoles or randomly in the cell cytoplasm, which might further damage the lysosomes and induce mitochondrial depolarization and cell apoptosis. Thus, the results from different studies should be compared with caution. In another study by another group, Wu et al. (2020) investigated the cytotoxicity and the transcriptomic profiles of PS particles in human Caco-2 cells (Wu et al., 2020). That work demonstrated that PS-MPLs reduced cell viability in a dose-dependent manner. The responsible genes were identified by Illumina RNA seq. The dominant pathways related to NF-κB, MAPK signalling, cytokine-cytokine receptor interaction, and toll-like receptors, which are involved in modulating cell inflammation and proliferation were strongly influenced. Moreover, the qPCR was applied to investigate the transcriptional level of five proliferation-related genes (Ras, ERK, MER, CDK4, Cyclin D1) and four inflammation-related genes (TRPV1, iNOS, IL-1β, IL-8), and the results were consistent with RNA-seq data. Lehner et al. (2020) (Lehner et al., 2020) designed a novel, three-dimensional in vitro intestinal model, consisting of the human intestinal epithelial cell lines Caco-2 and HT29-MTX-E12 as well as human blood monocyte-derived macrophages and dendritic cells, that is suitable for assessing the possible effects of ingested MPLs of different composition in automobile tyre wear and polyolefins in a range of sizes between 50 and 500 µm. MPL particles were exposed at concentrations of 823.5–1,380.0 µg/cm2 directly on the intestinal model’s surface. Cytotoxicity was investigated after 6, 24 and 48 h of exposure by measuring the release of lactate dehydrogenase. Inflammatory endpoints, including the cytokines IL-8, TNFα and IL-1β, as well as changes of the barrier integrity after exposure, were additionally monitored. The main results of this study were that the MPLs between 50 and 500 µm did not cause any significant cytotoxicity or release of (pro-)inflammatory cytokines and did not change the barrier integrity the points investigated, at any of the time. In another interesting study (Liu et al., 2020), using PS-MPLs (100 and 5,000 nm), the influence of the digestive process on intestinal toxicity was also studied by using the in vitro Caco-2 model (Liu et al., 2020). The toxicity of original and transformed MPLs was studied. Results showed that the digestive process did not alter the chemical constitution but formed a corona on the surface of MPLs. The 100 nm PS-MPLs showed higher intestinal toxicity than 5 μm PS-MPLs. Digestive treatment relieved cytotoxicity and transport function disorder of the Caco-2 monolayer at was induced by non-treated PS-MPLs. However, the in vitro digestive process increased the proinflammatory effects of PS-MPLs. The formation of a corona on the surface leads to a change in size, Zeta potential, and adsorbed compounds. Also, in the same line, Wang et al. (2021a) (Wang et al., 2021a) investigated the cytotoxicity of transformed PS-MPLs (presenting alterations in the surface of the particles, chemical composition, and changes of the surface charge) in hepatocytes. The results of that study revealed that 500 nm PS-MPLs, which were chemically transformed by simulated gastric fluid, exacerbated their toxicity on SMMC-7721 cells at 20 μg/ml after 24 h of treatment. The observed effects included morphological alteration of cells, membrane damage and increased cell apoptosis via oxidative stress. The authors hypothesised that these effects could be partially explained by the degradation and surface modifications of PS-MPLs. In conclusion, transformed PS-MPLs presented higher hepatic cytotoxicity after transformation by simulated gastric fluid.
The GIT uptake and effects of different MPLs of most common polymers (PE, PP, PET and PVC) were investigated by Stock et al. (Stock et al., 2021) using the Caco-2 cell line as the in vitro cell-line model and using the cytotoxicity of the human cell lines Caco-2, HepG2 and HepaRG in order to detect a possible impact on the organs which are first to come into contact with ingested particles: the intestine and the liver. The results of the study demonstrate that especially 1–4 μm PE-MPLs were transported through the intestinal epithelium and that intestinal exposure to MPLs is material- and size-dependent. However, it should be highlighted that only a high concentration far beyond realistic dietary exposure of consumers induced cytotoxic effects.
Although much less investigation has been undertaken in exploring the potential cytotoxicity of MNPLs in other tissues, several studies were carried out during recent years. Xu et al. (2019) (Xu et al., 2019) carried out a preliminary evaluation of the effects of PS-NPLs on human lung epithelial cells. The internalisation, cell viability, cell cycle, apoptosis, and associated gene transcription and protein expression were assessed on the human alveolar epithelial A549 cell line exposed to 25 and 70 nm PS-NPLs. Results showed that smaller particles were internalised more rapidly and efficiently into the cytoplasm of A549. PS-NPLs significantly affected the cell viability, induced significant up-regulation of pro-inflammatory cytokines such as IL-8, NF-κB, TNF-α, and pro-apoptotic proteins such as DR5, caspase-3, caspase-8, caspase-9, and cytochrome c, which revealed that it triggered a TNF-α-associated apoptosis pathway. That study suggests that exposure duration, diameter, and concentration are the key factors in the toxicological effects of PS-NPLs on alveolar epithelial cells. In another study by Dong et al. (2020) (Dong et al., 2020), the potential of in vitro pulmonary toxicity of PS-MPLs was assessed using normal human lung epithelial BEAS-2B cells. PS-MPLs with an average size of 1.72 ± 0.26 μm were used. Results revealed that PS-MPLs could cause cytotoxic and inflammatory effects in BEAS-2B cells by inducing reactive oxygen species (ROS) formation. PS-MPLs can decrease transepithelial electrical resistance by depleting zonula occludens (ZO) proteins. According to the authors, high concentrations of PS-MPLs produce the decreased α1-antitrypsin levels, increasing the risk of chronic obstructive pulmonary disease. Furthermore, low levels of PS-MPLs can disrupt the protective pulmonary barrier, and they may also increase the risk of lung disease. These findings indicate that PS-MPL inhalation may negatively influence human respiratory health. The potential toxicological damages of PS-MPLs in lung cell were also proposed by Goodman et al. (2021) (Goodman et al., 2021). In that study, human alveolar A549 cells were exposed to polystyrene microplastics (PS-MPLs) of 1 and 10 μm diameter. Both sizes caused a significant reduction in cell proliferation but exhibited little cytotoxicity. Despite these results, further assays revealed the metabolic activity decrease and a dramatic decrease in the proliferation rate of exposed cells. Phase-contrast imaging of live cells revealed at 72 h major changes in exposed cells morphology, as well as the uptake of 1 μm particles into the cells. Confocal fluorescent microscopy at 24 h of exposure confirmed the incorporation of 1 μm PS-MPLs.
Other studies have focussed on the potential genotoxic effects of PE on peripheral blood lymphocytes (Çobanoğlu et al., 2021) and the ROS effect on T98G and HeLa, cerebral and epithelial human cells, respectively (Schirinzi et al., 2017). In conclusion, the main toxicological effects that can be attributed to MNPLs particles are oxidative stress, cytotoxicity, and metabolic changes. However, the composition of the particles and the dose drive the toxicological properties, size, shape, surface functionalisation, or surface properties.
Pathways of Plastic Additive Toxicity
Plastic formulations are composed of polymer or polymer mixtures and plastic additives, including functionalising agents (plasticisers, impact modifiers, and flame retardants), stabilisers (antioxidants, UV-filters, thermal stabilisers), lubricants, fillers, and colourants. The additives are incorporated into the polymers during manufacture, but they are not chemically bound to them, as aforementioned. Due to this reason and their low molecular weight, plastic additives can leach the external medium along a concentration gradient (Tickner et al., 2001).
As mentioned before, when particle size decreases, the surface-active area in contact with the external medium will increase, facilitating the leach. If MNPLs accumulate in living organisms, they present a source of chemical plastic additives and polymer monomers to tissues and fluids.
Indeed, the impacts on human health of some selected groups of plastic additives such as polybrominated diphenyl ethers (PBDEs) (Maddela et al., 2020), bisphenol A (Kahn et al., 2020; Meli et al., 2020; Mustieles and Fernández, 2020a), and some phthalates (Sedha et al., 2021) have been extensively studied in the last 10 years, and relevant hazard and epidemiological information (Husøy et al., 2019) have been collected. These chemicals have been related to carcinogenicity, neurotoxicity (Mustieles and Fernández, 2020b; Li et al., 2020), obesity (Andújar et al., 2019), and endocrine disruption (Ma et al., 2019). Human studies showed that the concentrations of these additives in young children, a susceptible segment (Ma et al., 2019; Ouyang et al., 2020), are typically higher, indicating the need to decrease exposure to these compounds. Thus, a wide variety of substitution compounds have arisen in new plastic formulations, including emerging plasticisers, brominated flame retardants (BFR) and organophosphate esters (OPEs) in substituting PBDE legacies, or those used in new plastic materials such as bioplastics. Bioplastics derived from renewable biomass sources such as polylactic acid (PLA), polyhydroxyalkanoates (PHA), bio-derived polyethylene (Bio-PE), among many others, offer a greener solution for plastic production. However, very little information is currently available on their impact on living organisms, including humans. In this sense, Zimmermann et al. (2019) (Zimmermann et al., 2019) demonstrated that consumer products made of PLA induced strong baseline toxicity similar to PVC items. This demonstrates that this bio-based and biodegradable material, despite being marketed as better alternative, is not necessarily safer than conventional plastics. Despite the amount of research that has been devoted to assessing the potential impact of certain groups of plastic additives, such as some plasticisers and flame retardants, the information about the other groups continues to be scant. As can be seen in Table 4, this tendency has continued in recent years and most of the investigations have been focussed on plasticisers of the group of phthalates and bisphenols, and the new organ phosphate flame retardants. In addition to many toxicological studies showing the negative effects of certain phthalate congeners on human health, such as asthma, breast cancer, obesity, type II diabetes, and male infertility, the potential health damages by new substitution compounds, the phthalate transformation products, or even human metabolites of primary phthalates are open questions. In this regard, Sicińska (Sicińska, 2018), undertook an in vitro study of the induction of haemolysis and eryptosis in human erythrocytes caused by di-n-butyl phthalate (DBP), butyl benzyl phthalate (BBP), and their metabolites: mono-n-butyl phthalate (MBP) and mono-benzyl phthalate (MBzP). The main results of this study demonstrated that DBP and BBP possess higher haemolytic properties compared to their metabolites, but all compounds induced eryptosis.
TABLE 4. Examples toxicological assays of plastic additives using human cell-lines.
One of the most commonly used plasticisers is di (2-ethyl hexyl) phthalate (DEHP), which is metabolised to mono (2-ethylhexyl) phthalate (MEHP), and inhalation is an important exposure route for both. Their effects on the lungs include inflammation and alteration of postnatal maturation (alveolarisation). Recently, the cell progression, epithelial and mesenchymal markers were studied in alveolar epithelial cells A549, exposed to DEHP (1–100 μM) or MEHP (1–50 μM) for 24–72 h (Sedha et al., 2021). That experimental work showed that DEHP and MEHP altered the structure and migration of A549 cells and promoted the loss of the epithelial phenotype (Rafael-Vázquez et al., 2018). DEHP has been classified as an endocrine disruptor. Fang et al. (Fang et al., 2019a) investigated the mechanism of the embryotoxicity induced by DEHP in differentiated human embryonic stem cells (hESCs). It was demonstrated that DEHP exposure inhibited cell proliferation, promoted cell cycle arrest, and induced apoptosis through the PPARγ/PTEN/AKT signalling pathway, suggesting that DEHP exposure could cause reproductive or developmental toxicity in humans (Fang et al., 2019a).
Due to the endocrine disruption effects of different phthalates, currently, a series of substitutes have been proposed, such as tributyl o-acetylcitrate (ATBC), triethyl 2-acetylcitrate (ATEC), and trihexyl o-acetylcitrate (ATHC). To assess the safety of these new substitutes, the potential cytotoxicity of these compounds was screened in a human keratinocyte cell line HaCaT. The results demonstrated that ATEC showed similar levels of cytotoxicity as common phthalates, whereas ATBC and ATHC did not show significant cytotoxicity, even at very high doses (5 mg/ml) (Kim et al., 2019).
Another plasticiser with high usage in the past in polycarbonate plastic production is bisphenol-A (BPA). However, BPA exhibits toxicity, endocrine disruption, mutagenicity and carcinogenic effects in living organisms (Maffini et al., 2006; Rochester, 2013). For these reasons, in 2011, the plasticiser use was banned in baby bottles, cups, and other containers that are designed for use by children of 3 years old and younger, and in 2012 their use was also banned in water bottles. Currently, the toxicity of other alternative compounds of this group, such as bisphenol-F (BPF) and bisphenol-S (BPS), and 3,3′,5,5′-tetrabromobisphenol-A (TBBPA) that are also used as a flame retardant, are being investigated. Further, several studies have revealed that these compounds also presented toxicity in specific BPF and BPS (Fouyet et al., 2021).
Another group of plasticisers with increasing use are the adipates, this class of plasticisers has been much less studied but recent findings showed that bis (2-ethylhexyl) adipate (DEHA) and their metabolites can cause, even at low concentrations, endocrine-disrupting effects on the steroid hormone synthesis in H295R cell (Duan et al., 2020).
As aforementioned, flame retardants are another group of plastic additives that has been related to different toxicological effects in human health. In particular, some PBDEs are currently listed in Annex A of the Stockholm Convention on Persistent Organic Pollutants with specific exemptions for use regarding the recycling of articles that contain or may contain these chemicals. Thus, these compounds have been replaced by other groups, among them organophosphate-flame retardants. Currently, there are some gaps of information about their fate, behaviour and toxicity, but some compounds have been related to human accumulation (Chupeau et al., 2020) and health impairments (Wang et al., 2020a; Chupeau et al., 2020). Among them, there are five main toxicological effects described as reproductive and developmental toxicity (Liu et al., 2021b), neurotoxicity (Yan et al., 2021), hepatotoxicity (Zhu et al., 2021), cardiotoxicity (Reddam et al., 2019; Xiong et al., 2021) and endocrine disruption (Liu et al., 2019a). However, most of these studies were in vivo assays with animals, and some information of toxicity mechanisms from in vitro tests using human cell is still required.
Other chemicals that could leach from the plastic polymer matrix include antioxidants, UV stabilisers such as the group of benzophenones and plastic coating finishing compound such as perfluoroalkyl substances (PFASs) (Pelch et al., 2019), but with the exception of the PFASs group, the potential for adverse effects on human health of other plastic additives has been much less studied.
It is noteworthy that, currently, more than 5,300 plastic formulations are commercially available. More than 4,000 known chemicals are associated with plastic packaging alone, and to date most of these chemicals have not been appropriately studied.
Moreover, in general, the potential migration of additives from the plastics in use is the considered factor for the evaluation of their potential impact. In the case of MNPL contamination from different sources, the potential migration from MNPLs particles during long-term exposure should be assessed. Lixiviation should be evaluated from particles of the smallest range of sizes, with modified surface properties, considering the potential accumulation in tissues, and sometimes coming from plastic formulations that are no longer in use, with residues that are still in the environment.
Trojan-Horse Effects and Microbiome Effects
The increased surface area of MNPLs and their surface properties, such as hydrophobicity and adsorption, favour the accumulation onto their surfaces of other contaminants from surrounding environments, facilitating their potential transfer to living organisms, including humans. Llorca et al. (2018) (Llorca et al., 2018) assessed the adsorption and desorption behaviour of 18 perfluoroalkyl substances (PFASs) from the surrounding waters (freshwater and seawater) of Spain onto MNPLs surfaces of different polymers, including PE, PS, and polystyrene carboxylate (PS-COOH)). Then, the same group investigated the capacity for the transfer of polychlorinated biphenyls (PCBs), a group of persistent organic pollutants (POPs) to PE, PS, and PET MPL surfaces in water/sediment systems (Llorca et al., 2020b). In both studies, the adsorption/desorption of other contaminants onto MNPL surfaces was confirmed, mainly following the Freundlich isotherms. Medium polar compounds showed that the adsorption and posterior desorption increased. While highly non-polar compounds were sequestered onto the MNPL surfaces. Therefore, it should be considered how the interaction of contaminant/MPLs can affect the absolute bioavailability of these contaminants that are transferred to biota. Indeed, several works informed about the potential for the accumulation of environmental pollutants onto MPL surfaces (Lang et al., 2020; Atugoda et al., 2021; Fan et al., 2021; Puckowski et al., 2021; Torres et al., 2021). However, less research has been paid to assessing of whether certain tissue temperatures and physiological conditions can increase the potential transfer of other contaminants to humans. In this regard, the potential for POPs (DDT, PFOA, and DEHP) to desorb from MPLs of PVC and PE under simulated physiological conditions has been studied (Bakir et al., 2014). Desorption rates were enhanced up to 30 times compared to seawater at lower pH and higher temperatures, in conditions simulating the physiological conditions of warm organisms, 38°C, pH 4. Therefore, this pathway of human exposure needs to be addressed. In general, the toxicological interaction between MNPLs and other contaminants has been assessed using in vivo studies, the majority of which were focused on the effect on model organisms exposed to a mixture of chemicals or an individual chemical but in both cases at a single concentration. Therefore, dose-response curves were not provided, and a complete description of these interactions was not assessed.
Several factors rule the modulation of the toxicity of MNPLs and other contaminant toxicity such as physicochemical properties of MNPLs and contaminant mixtures, and therefore, their adsorption/desorption behaviour, and that of the selected model organisms. Other essential factors are the size and shape of MNPLs, and even with respect to the colour for some test organisms (Bhagat et al., 2021).
On the other hand, MNPL surfaces are colonised by microbes. In the natural environment, for example, biofilms can be established on plastic particle surfaces within a period of 7 days. Thus, the potential transfer of pathogens via microplastics that are inhaled or ingested should also be investigated.
Finally, another potential effect that should be closely studied is the MNPLs potential to disrupt the human microbiome. The chronic ingestion of MNPLs could impact on the natural community and abundance pattern of the gut microbiota. This so-called dysbiosis could be associated to mechanical disruption within the GIT, the ingestion of foreign and potentially pathogenic bacteria, as well as chemicals, which make-up or adhere to MNPLs. The impact on the microbiome can result in the immune system response and trigger chronic diseases, promote infections, and alter the gut microbiota. For example, lipopolysaccharides (LPS) are able to adjust the hydrophobicity in their cell walls by changing their composition in the outer membrane, to better interact with other hydrophobic substances (Krasowska and Sigler, 2014; Rogers et al., 2020). LPS also can effect gut microbiota so changes on LPS will direct affect microbiota (Di Lorenzo et al., 2019). In this regard, Fackelmann and Sommer (Fackelmann and Sommer, 2019) have summarised microplastics-induced gut dysbiosis effects on host health, and some few pioneering studies have been carried out during the last years, studying responses on mice (Lu et al., 2018) and aquatic organisms (Evariste et al., 2019), but this is another human open question for human health implications.
Current Needs and Future Research Trends
The study of the potential impacts of MNPLs in human health is in an early stage of development. Firstly, the analytical approaches to assess them are limited by the size of the particles, and there are only a few approaches able to assess the particles at the nano-range. Thus, human exposure assessment continues to be a challenge. Moreover, it is expected that small-sized particles are those with a superior rate of up-take and translocation through tissues, but this concordance is not always observed, and it should be better studied. On the other hand, the interaction of MNPLs to lipopolysaccharides needs to be further investigated since the effects of LPS on the gut plays an important role and their release into the blood will damage the endothelium. This needs to be examined because the effects of LPS on the gut is important and the release of LPS into the blood will damage the endothelium.
Particle weathering is also a parameter that should be considered when assessing toxicity of MNPLs- Physical damages caused by particles include the disruption of the immune system, the increase of barriers permeability, leading to oxidative stress and inflammation, favouring the uptake of smallest particles and translocation through tissues. In addition, their higher surface area favours the long-term leach of plastic additives or the transfer to the biota of other contaminants and maybe pathogens. Besides, the potential disturbance of the human microbiome can be another cause of human health damage almost not studied until now.
The main routes of human exposure have already been established, with ingestion and inhalation being the main routes. Nevertheless, the occurrence of MNPLs in complete diets has not been established to date. Therefore, the total contribution of MNPLs to human exposure of exogenous particles through ingestion is still unknown. Moreover, incidental sources of exposure (MNPLs from food packaging, ingestion from PPCs, among others, have not been well determined). Inhalation is another relevant source of MNPL exposure, but relatively few studies have considered indoor exposure.
There is a need for more comprehensive toxicological studies, including particulate physical damages and the toxicity of additives. In this regard, the information about plastic additive toxicity continues being limited to targeted groups of compounds such as phthalates. In addition, the toxicity of MNPLs is also modulated by their size, shape, and studies comparing these variables are necessary.
Finally, to assess the impacts of MNPLs contamination and their potential damages to human health, cohort studies focussed on this emerging type of contamination are required.
Data Availability Statement
The original contributions presented in the study are included in the article/Supplementary Material, further inquiries can be directed to the corresponding authors.
Author Contributions
ML has contributed to the design and writing of diffeerent sections of the document. MF concieled the idea of the manuscript, designed it and wrote different sections.
Funding
This work was supported by PLAS-MED and DEGBIOPLAST projects (CTM 2017–-89701-C3-1-R and RTI 2018-097860-J-I00) from the Spanish Ministry of Science and Innovation.
Conflict of Interest
The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.
Publisher’s Note
All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.
References
Abbasi, S., Keshavarzi, B., Moore, F., Turner, A., Kelly, F. J., Dominguez, A. O., et al. (2019). Distribution and Potential Health Impacts of Microplastics and Microrubbers in Air and Street Dusts from Asaluyeh County, Iran. Environ. Pollut. 244, 153–164. doi:10.1016/j.envpol.2018.10.039
Abidli, S., Akkari, N., Lahbib, Y., Trigui, N., and Menif, E. (2021). First Evaluation of Microplastics in Two Commercial Fish Species from the Lagoons of Bizerte and Ghar El Melh (Northern Tunisia). Reg. Stud. Mar. Sci. 41, 101581. doi:10.1016/j.rsma.2020.101581
Akhbarizadeh, R., Dobaradaran, S., Nabipour, I., Tajbakhsh, S., Darabi, A. H., and Spitz, J. (2020). Abundance, Composition, and Potential Intake of Microplastics in Canned Fish. Mar. Pollut. Bull. 160, 111633. doi:10.1016/j.marpolbul.2020.111633
Akhbarizadeh, R., Dobaradaran, S., Schmidt, T. C., Nabipour, I., and Spitz, J. (2020). Worldwide Bottled Water Occurrence of Emerging Contaminants: A Review of the Recent Scientific Literature. J. Hazard. Mater. 392, 122271. doi:10.1016/j.jhazmat.2020.122271
Akhbarizadeh, R., Moore, F., and Keshavarzi, B. (2019). Investigating Microplastics Bioaccumulation and Biomagnification in Seafood from the Persian Gulf: a Threat to Human Health. Food Additives & Contaminants: A 36 (11), 1696–1708. doi:10.1080/19440049.2019.1649473
Al-Salem, A. M., Saquib, Q., Siddiqui, M. A., Ahmad, J., Wahab, R., and Al-Khedhairy, A. A. (2019). Organophosphorus Flame Retardant (Tricresyl Phosphate) Trigger Apoptosis in HepG2 Cells: Transcriptomic Evidence on Activation of Human Cancer Pathways. Chemosphere 237, 124519. doi:10.1016/j.chemosphere.2019.124519
Allen, S., Allen, D., Phoenix, V. R., Le Roux, G., Durántez Jiménez, P., Simonneau, A., et al. (2019). Atmospheric Transport and Deposition of Microplastics in a Remote Mountain Catchment. Nat. Geosci. 12 (5), 339–344. doi:10.1038/s41561-019-0335-5
Andrady, A. L. (2011). Microplastics in the marine Environment. Mar. Pollut. Bull. 62 (8), 1596–1605. doi:10.1016/j.marpolbul.2011.05.030
Andrady, A. L., and Neal, M. A. (2009). Applications and Societal Benefits of Plastics. Phil. Trans. R. Soc. B. 364 (1526), 1977–1984. doi:10.1098/rstb.2008.0304
Andújar, N., Gálvez-Ontiveros, Y., Zafra-Gómez, A., Rodrigo, L., Álvarez-Cubero, M. J., Aguilera, M., et al. (2019). Bisphenol A Analogues in Food and Their Hormonal and Obesogenic Effects: A Review. Nutrients 11 (9). doi:10.3390/nu11092136
Atugoda, T., Vithanage, M., Wijesekara, H., Bolan, N., Sarmah, A. K., Bank, M. S., et al. (2021), Interactions between Microplastics, Pharmaceuticals and Personal Care Products: Implications for Vector Transport. Environ. Int. 149. 106367. doi:10.1016/j.envint.2020.106367
Baechler, B. R., . Granek, E. F., Mazzone, S. J., Nielsen-Pincus, M., and Brander, S. M. (2020). Microplastic Exposure by Razor Clam Recreational Harvester-Consumers along a Sparsely Populated Coastline. Front. Mar. Sci. 7. doi:10.3389/fmars.2020.588481
Bahrina, I., Syafei, A., Satoto, R., Jiang, J.-J., Nurasrin, N., Assomadi, A., et al. (2020). An Occupant-Based Overview of Microplastics in Indoor Environments in the City of Surabaya, Indonesia. J. Ecol. Eng. 21 (8), 236–242. doi:10.12911/22998993/126876
Bakir, A., Rowland, S. J., and Thompson, R. C. (2014). Enhanced Desorption of Persistent Organic Pollutants from Microplastics under Simulated Physiological Conditions. Environ. Pollut. 185, 16–23. doi:10.1016/j.envpol.2013.10.007
Barboza, L. G. A., Lopes, C., Oliveira, P., Bessa, F., Otero, V., Henriques, B., et al. (2020). Microplastics in Wild Fish from North East Atlantic Ocean and its Potential for Causing Neurotoxic Effects, Lipid Oxidative Damage, and Human Health Risks Associated with Ingestion Exposure. Sci. Total Environ. 717, 134625. doi:10.1016/j.scitotenv.2019.134625
Besseling, E., Wegner, A., Foekema, E. M., van den Heuvel-Greve, M. J., and Koelmans, A. A. (2013). Effects of Microplastic on Fitness and PCB Bioaccumulation by the Lugworm Arenicola marina (L.). Environ. Sci. Technol. 47 (1), 593–600. doi:10.1021/es302763x
Bhagat, J., Nishimura, N., and Shimada, Y. (2021). Toxicological Interactions of Microplastics/nanoplastics and Environmental Contaminants: Current Knowledge and Future Perspectives. J. Hazard. Mater. 405, 123913. doi:10.1016/j.jhazmat.2020.123913
Bonello, G., Varrella, P., and Pane, L. (2018). First Evaluation of Microplastic Content in Benthic Filter-Feeders of the Gulf of La Spezia (Ligurian Sea). J. Aquat. Food Product. Tech. 27 (3), 284–291. doi:10.1080/10498850.2018.1427820
Bouwmeester, H., Hollman, P. C. H., and Peters, R. J. B. (2015). Potential Health Impact of Environmentally Released Micro- and Nanoplastics in the Human Food Production Chain: Experiences from Nanotoxicology. Environ. Sci. Technol. 49 (15), 8932–8947. doi:10.1021/acs.est.5b01090
Campanale, C., Massarelli, fnm., Savino, fnm., Locaputo, fnm., and Uricchio, fnm. (2020). A Detailed Review Study on Potential Effects of Microplastics and Additives of Concern on Human Health. Ijerph 17 (4), 1212. doi:10.3390/ijerph17041212
Catarino, A. I., Macchia, V., Sanderson, W. G., Thompson, R. C., and Henry, T. B. (2018). Low Levels of Microplastics (MP) in Wild Mussels Indicate that MP Ingestion by Humans Is Minimal Compared to Exposure via Household Fibres Fallout during a Meal. Environ. Pollut. 237, 675–684. doi:10.1016/j.envpol.2018.02.069
Cho, S. A., Cho, W. B., Kim, S. B., Chung, J. H., Kim, H. J., et al. (2019). Identification of Microplastics in Sea Salts by Raman Microscopy and FT-IR Microscopy. Anal. Sci. Tech. 32 (6), 243–251.
Chupeau, Z., Bonvallot, N., Mercier, F., Le Bot, B., Chevrier, C., and Glorennec, P. (2020). Organophosphorus Flame Retardants: A Global Review of Indoor Contamination and Human Exposure in Europe and Epidemiological Evidence. Int. J. Environ. Res. Public Health 17 (18), 1–24. doi:10.3390/ijerph17186713
Çobanoğlu, H., Belivermiş, M., Sıkdokur, E., Kılıç, Ö., and Çayır, A. (2021). Genotoxic and Cytotoxic Effects of Polyethylene Microplastics on Human Peripheral Blood Lymphocytes. Chemosphere 272, 129805. doi:10.1016/j.chemosphere.2021.129805
Cole, M., Lindeque, P., Halsband, C., and Galloway, T. S. (2011). Microplastics as Contaminants in the marine Environment: A Review. Mar. Pollut. Bull. 62 (12), 2588–2597. doi:10.1016/j.marpolbul.2011.09.025
Coppola, F., Bessa, A., Henriques, B., Russo, T., Soares, A. M. V. M., Figueira, E., et al. (2020). Oxidative Stress, Metabolic and Histopathological Alterations in Mussels Exposed to Remediated Seawater by GO-PEI after Contamination with Mercury. Comp. Biochem. Physiol. A. Mol. Integr. Physiol. 243, 110674. doi:10.1016/j.cbpa.2020.110674
Cox, K. D., Covernton, G. A., Davies, H. L., Dower, J. F., Juanes, F., and Dudas, S. E. (2019). Human Consumption of Microplastics. Environ. Sci. Technol. 53 (12), 7068–7074. doi:10.1021/acs.est.9b01517
Daniel, D. B., Ashraf, P. M., and Thomas, S. N. (2020). Microplastics in the Edible and Inedible Tissues of Pelagic Fishes Sold for Human Consumption in Kerala, India. Environ. Pollut. 266, 115365. doi:10.1016/j.envpol.2020.115365
Di Lorenzo, F., De Castro, C., Silipo, A., and Molinaro, A. (2019). Lipopolysaccharide Structures of Gram-Negative Populations in the Gut Microbiota and Effects on Host Interactions. FEMS Microbiol. Rev. 43 (3), 257–272. doi:10.1093/femsre/fuz002
Ding, J., Sun, C., He, C., Li, J., Ju, P., and Li, F. (2021). Microplastics in Four Bivalve Species and Basis for Using Bivalves as Bioindicators of Microplastic Pollution. Science of the Total Environment, 782. doi:10.1016/j.scitotenv.2021.146830
Dong, C. D., Chen, C. W., Chen, Y. C., Chen, H. H., Lee, J. S., and Lin, C. H. (2020). Polystyrene Microplastic Particles: In Vitro Pulmonary Toxicity Assessment. J. Hazard. Mater. 385, 121575. doi:10.1016/j.jhazmat.2019.121575
Dowarah, K., Patchaiyappan, A., Thirunavukkarasu, C., Jayakumar, S., and Devipriya, S. P. (2020). Quantification of Microplastics Using Nile Red in Two Bivalve Species Perna Viridis and Meretrix Meretrix from Three Estuaries in Pondicherry, India and Microplastic Uptake by Local Communities through Bivalve Diet. Mar. Pollut. Bull. 153, 110982. doi:10.1016/j.marpolbul.2020.110982
Dris, R., Gasperi, J., Saad, M., Mirande, C., and Tassin, B. (2016). Synthetic Fibers in Atmospheric Fallout: A Source of Microplastics in the Environment. Mar. Pollut. Bull. 104 (1-2), 290–293. doi:10.1016/j.marpolbul.2016.01.006
Dris, R., Gasperi, J., Mirande, C., Mandin, C., Guerrouache, M., Langlois, V., et al. (2017). A First Overview of Textile Fibers, Including Microplastics, in Indoor and Outdoor Environments. Environ. Pollut. 221, 453–458. doi:10.1016/j.envpol.2016.12.013
Dris, R., Gasperi, J., Rocher, V., Saad, M., Renault, N., and Tassin, B. (2015). Microplastic Contamination in an Urban Area: A Case Study in Greater Paris. Environ. Chem. 12 (5), 592–599. doi:10.1071/en14167
Du, F., Cai, H., Zhang, Q., Chen, Q., and Shi, H. (2020). Microplastics in Take-Out Food Containers. J. Hazard. Mater. 399, 122969. doi:10.1016/j.jhazmat.2020.122969
Duan, C., Fang, Y., Sun, J., Li, Z., Wang, Q., Bai, J., et al. (2020). Effects of Fast Food Packaging Plasticizers and Their Metabolites on Steroid Hormone Synthesis in H295R Cells, Sci. Total Environ., 726. 138500. doi:10.1016/j.scitotenv.2020.138500
Eerkes-Medrano, D., Thompson, R. C., and Aldridge, D. C. (2015). Microplastics in Freshwater Systems: A Review of the Emerging Threats, Identification of Knowledge Gaps and Prioritisation of Research Needs. Water Res. 75, 63–82. doi:10.1016/j.watres.2015.02.012
Efsa (2016). Presence of Microplastics and Nanoplastics in Food, with Particular Focus on Seafood. EFSA J. 14 (6).
Evariste, L., Barret, M., Mottier, A., Mouchet, F., Gauthier, L., and Pinelli, E. (2019). Gut Microbiota of Aquatic Organisms: A Key Endpoint for Ecotoxicological Studies. Environ. Pollut. 248, 989–999. doi:10.1016/j.envpol.2019.02.101
Fackelmann, G., and Sommer, S. (2019). Microplastics and the Gut Microbiome: How Chronically Exposed Species May Suffer from Gut Dysbiosis. Mar. Pollut. Bull. 143, 193–203. doi:10.1016/j.marpolbul.2019.04.030
Fan, X., Gan, R., Liu, J., Xie, Y., Xu, D., Xiang, Y., et al. (2021). Adsorption and Desorption Behaviors of Antibiotics by Tire Wear Particles and Polyethylene Microplastics with or without Aging Processes, Sci. Total Environ., 771. 145451. doi:10.1016/j.scitotenv.2021.145451
Fang, C., Zheng, R., Chen, H., Hong, F., Lin, L., Lin, H., et al. (2019). Comparison of Microplastic Contamination in Fish and Bivalves from Two Major Cities in Fujian Province, China and the Implications for Human Health. Aquaculture 512, 734322. doi:10.1016/j.aquaculture.2019.734322
Fang, H., Fang, W., Cao, H., Luo, S., Cong, J., Liu, S., et al. (2019). Di-(2-ethylhexyl)-phthalate Induces Apoptosis via the PPARγ/PTEN/AKT Pathway in Differentiated Human Embryonic Stem Cells. Food Chem. Toxicol. 131, 110552. doi:10.1016/j.fct.2019.05.060
Fouyet, S., Olivier, E., Leproux, P., Dutot, M., and Rat, P. (2021). Bisphenol A, Bisphenol F, and Bisphenol S: The Bad and the Ugly. Where Is the Good. Life (Basel) 11 (4). doi:10.3390/life11040314
Galloway, T. S. (2015). Micro- and Nano-Plastics and Human Health. Mar. Anthropogenic Litter, 343–366. doi:10.1007/978-3-319-16510-3_13
Ghosh, G. C., AkterIslam, S. M., Rashidul, M., Habib, A., Chakraborty, T. K., and Zaman, S. (2021). Microplastics Contamination in Commercial marine Fish from the Bay of Bengal. Reg. Stud. Mar. Sci. 44, 101728. doi:10.1016/j.rsma.2021.101728
Goodman, K. E., Pineles, L., Magder, L. S., O'Hara, L. M., Pineles, B. L., Nadimpalli, G., et al. (2021). Exposure of Human Lung Cells to Polystyrene Microplastics Significantly Retards Cell Proliferation and Triggers Morphological Changes. Chem. Res. Toxicol. 34 (4), 1069–1081. doi:10.1021/acs.chemrestox.0c00486-
Hadibarata, T., Sathishkumar, P., Prasetia, H., Hikmat, , Pusfitasari, E. D., et al. (2021). Microplastic Contamination in the Skipjack Tuna (Euthynnus Affinis) Collected from Southern Coast of Java, Indonesia. Chemosphere 276, 130185. doi:10.1016/j.chemosphere.2021.130185
Hermabessiere, L., Dehaut, A., Paul-Pont, I., Lacroix, C., Jezequel, R., Soudant, P., et al. (2017). Occurrence and Effects of Plastic Additives on marine Environments and Organisms: A Review. Chemosphere 182, 781–793. doi:10.1016/j.chemosphere.2017.05.096
Hermabessiere, L., Paul-Pont, I., Cassone, A.-L., Himber, C., Receveur, J., Jezequel, R., et al. (2019). Microplastic Contamination and Pollutant Levels in Mussels and Cockles Collected along the Channel Coasts. Environ. Pollut. 250, 807–819. doi:10.1016/j.envpol.2019.04.051
Hollóczki, O., and Gehrke, S. (2019). Nanoplastics Can Change the Secondary Structure of Proteins. Sci. Rep. 9 (1), 16013. doi:10.1038/s41598-019-52495-w
Hollóczki, O., and Gehrke, S. (2020). Can Nanoplastics Alter Cell Membranes. ChemPhysChem 21 (1), 3. doi:10.1002/cphc.201901156
Horton, A. A., Walton, A., Spurgeon, D. J., Lahive, E., and Svendsen, C. (2017). Microplastics in Freshwater and Terrestrial Environments: Evaluating the Current Understanding to Identify the Knowledge Gaps and Future Research Priorities. Sci. Total Environ. 586, 127–141. doi:10.1016/j.scitotenv.2017.01.190
Huang, Y., Qing, X., Wang, W., Han, G., and Wang, J. (2020). Mini-review on Current Studies of Airborne Microplastics: Analytical Methods, Occurrence, Sources, Fate and Potential Risk to Human Beings. Trac Trends Anal. Chem. 125, 115821. doi:10.1016/j.trac.2020.115821
Huerta Lwanga, E., Mendoza Vega, J., Ku Quej, V., Chi, J. d. l. A., Sanchez del Cid, L., Chi, C., et al. (2017). Field Evidence for Transfer of Plastic Debris along a Terrestrial Food Chain. Sci. Rep. 7 (1), 14071. doi:10.1038/s41598-017-14588-2
Husøy, T., Andreassen, M., Hjertholm, H., Carlsen, M. H., Norberg, N., Sprong, C., et al. (2019). The Norwegian Biomonitoring Study from the EU Project EuroMix: Levels of Phenols and Phthalates in 24-hour Urine Samples and Exposure Sources from Food and Personal Care Products. Environ. Int. 132, 105103. doi:10.1016/j.envint.2019.105103
Hwang, J., Choi, D., Han, S., Jung, S. Y., Choi, J., and Hong, J. (2020). Potential Toxicity of Polystyrene Microplastic Particles. Sci. Rep. 10 (1), 7391. doi:10.1038/s41598-020-64464-9
Ivar Do Sul, J. A., and Costa, M. F. (2014). The Present and Future of Microplastic Pollution in the marine Environment. Environ. Pollut. 185, 352–364. doi:10.1016/j.envpol.2013.10.036
Kahn, L. G., Philippat, C., Nakayama, S. F., Slama, R., and Trasande, L. (2020). Endocrine-disrupting Chemicals: Implications for Human Health. Lancet Diabetes Endocrinol. 8 (8), 703–718. doi:10.1016/s2213-8587(20)30129-7
Kambia, N., Séverin, I., Farce, A., Dahbi, L., Dine, T., Moreau, E., et al. (2021). Comparative Effects of Di-(2-ethylhexyl)phthalate and Di-(2-ethylhexyl)terephthalate Metabolites on Thyroid Receptors: In Vitro and In Silico Studies. Metabolites 11 (2), 1–18. doi:10.3390/metabo11020094
Karami, A., Golieskardi, A., Choo, C. K., Larat, V., Karbalaei, S., and Salamatinia, B. (2018). Microplastic and Mesoplastic Contamination in Canned Sardines and Sprats. Sci. Total Environ. 612, 1380–1386. doi:10.1016/j.scitotenv.2017.09.005
Karlsson, T. M., Vethaak, A. D., Almroth, B. C., Ariese, F., van Velzen, M., Hassellöv, M., et al. (2017). Screening for Microplastics in Sediment, Water, marine Invertebrates and Fish: Method Development and Microplastic Accumulation. Mar. Pollut. Bull. 122 (1), 403–408. doi:10.1016/j.marpolbul.2017.06.081
Kedzierski, M., Lechat, B., Sire, O., Le Maguer, G., Le Tilly, V., and Bruzaud, S. (2020). Microplastic Contamination of Packaged Meat: Occurrence and Associated Risks. Food Packaging and Shelf Life 24, 100489. doi:10.1016/j.fpsl.2020.100489
Kim, H., Nam, K., Oh, S., Son, S., Jeon, D., Gye, M. C., et al. (2019). Toxicological Assessment of Phthalates and Their Alternatives Using Human Keratinocytes. Environ. Res. 175, 316–322. doi:10.1016/j.envres.2019.05.007
Knobloch, E., Ruffell, H., Aves, A., Pantos, O., Gaw, S., and Revell, L. E. (2021). Comparison of Deposition Sampling Methods to Collect Airborne Microplastics in Christchurch, New Zealand. Water Air Soil Pollut. 232 (4), 133. doi:10.1007/s11270-021-05080-9
Kosuth, M., Mason, S. A., and Wattenberg, E. V. (2018). Anthropogenic Contamination of Tap Water, Beer, and Sea Salt. PLoS ONE 13 (4), e0194970. doi:10.1371/journal.pone.0194970
Krasowska, A., and Sigler, K. (2014). How Microorganisms Use Hydrophobicity and what Does This Mean for Human Needs. Front Cel Infect Microbiol. 4 (112), 112. doi:10.3389/fcimb.2014.00112
Krivoshiev, B. V., Beemster, G. T. S., Sprangers, K., Cuypers, B., Laukens, K., Blust, R., et al. (2018). Toxicogenomics of the Flame Retardant Tris (2-butoxyethyl) Phosphate in HepG2 Cells Using RNA-Seq. Toxicol. Vitro 46, 178–188. doi:10.1016/j.tiv.2017.10.011
Laborda, F., Trujillo, C., and Lobinski, R. (2021). Analysis of Microplastics in Consumer Products by Single Particle-Inductively Coupled Plasma Mass Spectrometry Using the Carbon-13 Isotope. Talanta 221, 121486. doi:10.1016/j.talanta.2020.121486
Lang, M., Yu, X., Liu, J., Xia, T., Wang, T., Jia, H., et al. (2020). Fenton Aging Significantly Affects the Heavy Metal Adsorption Capacity of Polystyrene Microplastics. Sci. Total Environ. 722, 137762. doi:10.1016/j.scitotenv.2020.137762
Lau, O. W., and Wong, S. K. (2000). Contamination in Food from Packaging Material. J. Chromatogr. A. 882 (1), 255–270. doi:10.1016/s0021-9673(00)00356-3
Lehner, R., Wohlleben, W., Septiadi, D., Landsiedel, R., Petri-Fink, A., and Rothen-Rutishauser, B. (2020). A Novel 3D Intestine Barrier Model to Study the Immune Response upon Exposure to Microplastics. Arch. Toxicol. 94 (7), 2463–2479. doi:10.1007/s00204-020-02750-1
Li, F., Yang, F., Li, D. K., Tian, Y., Miao, M., Zhang, Y., et al. (2020). Prenatal Bisphenol A Exposure, Fetal Thyroid Hormones and Neurobehavioral Development in Children at 2 and 4 years: A Prospective Cohort Study. Sci. Total Environ. 722, 137887. doi:10.1016/j.scitotenv.2020.137887
Li, J., Yang, D., Li, L., Jabeen, K., and Shi, H. (2015). Microplastics in Commercial Bivalves from China. Environ. Pollut. 207, 190–195. doi:10.1016/j.envpol.2015.09.018
Li, Y., Chen, G., Xu, K., Huang, K., and Wang, J. (2021). Microplastics Environmental Effect and Risk Assessment on the Aquaculture Systems from south China. Int. J. Environ. Res. Public Health 18 (4), 1–14. doi:10.3390/ijerph18041869
Liu, C., Li, J., Zhang, Y., Wang, L., Deng, J., Gao, Y., et al. (2019). Widespread Distribution of PET and PC Microplastics in Dust in Urban China and Their Estimated Human Exposure. Environ. Int. 128, 116–124. doi:10.1016/j.envint.2019.04.024
Liu, K., Wu, T., Wang, X., Song, Z., Zong, C., Wei, N., et al. (2019). Consistent Transport of Terrestrial Microplastics to the Ocean through Atmosphere. Environ. Sci. Technol. 53 (18), 10612–10619. doi:10.1021/acs.est.9b03427
Liu, L., Xu, K., Zhang, B., Ye, Y., Zhang, Q., and Jiang, W. (2021). Cellular Internalization and Release of Polystyrene Microplastics and Nanoplastics. Sci. Total Environ. 779, 146523. doi:10.1016/j.scitotenv.2021.146523
Liu, S., Wu, X., Gu, W., Yu, J., and Wu, B. (2020). Influence of the Digestive Process on Intestinal Toxicity of Polystyrene Microplastics as Determined by In Vitro Caco-2 Models. Chemosphere 256, 127204. doi:10.1016/j.chemosphere.2020.127204
Liu, W., Luo, D., Xia, W., Tao, Y., Wang, L., Yu, M., et al. (2021). Prenatal Exposure to Halogenated, Aryl, and Alkyl Organophosphate Esters and Child Neurodevelopment at Two Years of Age. J. Hazard. Mater. 408, 124856. doi:10.1016/j.jhazmat.2020.124856
Liu, X., Cai, Y., Wang, Y., Xu, S., Ji, K., and Choi, K. (2019). Effects of Tris(1,3-Dichloro-2-Propyl) Phosphate (TDCPP) and Triphenyl Phosphate (TPP) on Sex-dependent Alterations of Thyroid Hormones in Adult Zebrafish. Ecotoxicology Environ. Saf. 170, 25–32. doi:10.1016/j.ecoenv.2018.11.058
Llorca, M., Ábalos, M., Vega-Herrera, A., Adrados, M. A., Abad, E., and Farré, M. (2020). Adsorption and Desorption Behaviour of Polychlorinated Biphenyls onto Microplastics' Surfaces in Water/sediment Systems. Toxics 8 (3). doi:10.3390/toxics8030059
Llorca, M., Álvarez-Muñoz, D., Ábalos, M., Rodríguez-Mozaz, S., H.M.L.M. Santos, L., and M. León, V. (2020). Microplastics in Mediterranean Coastal Area: Toxicity and Impact for the Environment and Human Health. Trends Environ. Anal. Chem. 27, e00090. doi:10.1016/j.teac.2020.e00090
Llorca, M., Schirinzi, G., Martínez, M., Barceló, D., and Farré, M. (2018). Adsorption of Perfluoroalkyl Substances on Microplastics under Environmental Conditions. Environ. Pollut. 235, 680–691. doi:10.1016/j.envpol.2017.12.075
Lu, L., Wan, Z., Luo, T., Fu, Z., and Jin, Y. (2018). Polystyrene Microplastics Induce Gut Microbiota Dysbiosis and Hepatic Lipid Metabolism Disorder in Mice. Sci. Total Environ. 631-632, 449–458. doi:10.1016/j.scitotenv.2018.03.051
Lunov, O., Syrovets, T., Loos, C., Beil, J., Delacher, M., Tron, K., et al. (2011). Differential Uptake of Functionalized Polystyrene Nanoparticles by Human Macrophages and a Monocytic Cell Line. ACS nano 5, 1657–1669. doi:10.1021/nn2000756
Ma, Y., Liu, H., Wu, J., Yuan, L., Wang, Y., Du, X., et al. (2019). The Adverse Health Effects of Bisphenol A and Related Toxicity Mechanisms. Environ. Res. 176, 108575. doi:10.1016/j.envres.2019.108575
Maddela, N. R., Venkateswarlu, K., Kakarla, D., and Megharaj, M. (2020). Inevitable Human Exposure to Emissions of Polybrominated Diphenyl Ethers: A Perspective on Potential Health Risks. Environ. Pollut. 266, 115240. doi:10.1016/j.envpol.2020.115240
Maffini, M. V., Rubin, B. S., Sonnenschein, C., and Soto, A. M. (2006). Endocrine Disruptors and Reproductive Health: The Case of Bisphenol-A. Mol. Cell Endocrinol. 254-255, 179–186. doi:10.1016/j.mce.2006.04.033
Makhdoumi, P., Amin, A. A., Karimi, H., Pirsaheb, M., Kim, H., and Hossini, H. (2021). Occurrence of Microplastic Particles in the Most Popular Iranian Bottled mineral Water Brands and an Assessment of Human Exposure. J. Water Process Eng. 39. doi:10.1016/j.jwpe.2020.101708
Marques, F., Vale, C., Rudnitskaya, A., Moreirinha, C., Costa, S. T., and Botelho, M. J. (2021). Major Characteristics of Microplastics in Mussels from the Portuguese Coast. Environ. Res. 197, 110993. doi:10.1016/j.envres.2021.110993
Mason, S. A., Welch, V. G., and Neratko, J. (2018). Synthetic Polymer Contamination in Bottled Water. Front. Chem. 6, 407. doi:10.3389/fchem.2018.00407
Meli, R., Monnolo, A., Annunziata, C., Pirozzi, C., and Ferrante, M. C. (2020). Oxidative Stress and BPA Toxicity: An Antioxidant Approach for Male and Female Reproductive Dysfunction. Antioxidants (Basel) 9 (5). doi:10.3390/antiox9050405
Mintenig, S. M., Löder, M. G. J., Primpke, S., and Gerdts, G. (2019). Low Numbers of Microplastics Detected in Drinking Water from Ground Water Sources. Sci. Total Environ. 648, 631–635. doi:10.1016/j.scitotenv.2018.08.178
Mustieles, V., and Fernández, M. F. (2020). Bisphenol A Shapes Children's Brain and Behavior: Towards an Integrated Neurotoxicity Assessment Including Human Data. Environ. Health 19 (1), 66. doi:10.1186/s12940-020-00620-y
Mustieles, V., and Fernández, M. F. (2020). Bisphenol A Shapes Children's Brain and Behavior: towards an Integrated Neurotoxicity Assessment Including Human Data. Environ. Health 19 (1), 66. doi:10.1186/s12940-020-00620-y
Naji, A., Nuri, M., and Vethaak, A. D. (2018). Microplastics Contamination in Molluscs from the Northern Part of the Persian Gulf. Environ. Pollut. 235, 113–120. doi:10.1016/j.envpol.2017.12.046
Nalbone, L., Cincotta, F., Giarratana, F., Ziino, G., and Panebianco, A. (2021). Microplastics in Fresh and Processed Mussels Sampled from Fish Shops and Large Retail Chains in Italy. Food Control 125, 108003. doi:10.1016/j.foodcont.2021.108003
Oliveri Conti, G., Ferrante, M., Banni, M., Favara, C., Nicolosi, I., Cristaldi, A., et al. (2020). Micro- and Nano-Plastics in Edible Fruit and Vegetables. The First Diet Risks Assessment for the General Population. Environ. Res. 187, 109677. doi:10.1016/j.envres.2020.109677
Ouyang, F., Zhang, G. H., Du, K., Shen, L., Ma, R., Wang, X., et al. (2020). Maternal Prenatal Urinary Bisphenol A Level and Child Cardio-Metabolic Risk Factors: A Prospective Cohort Study. Environ. Pollut. 265, 115008. doi:10.1016/j.envpol.2020.115008
Oßmann, B. E. (2021). Microplastics in Drinking Water? Present State of Knowledge and Open Questions. Curr. Opin. Food Sci. 41, 44–51. doi:10.1016/j.cofs.2021.02.011
Pelch, K. E., Reade, A., Wolffe, T. A. M., and Kwiatkowski, C. F. (2019). PFAS Health Effects Database: Protocol for a Systematic Evidence Map. Environ. Int. 130, 104851. doi:10.1016/j.envint.2019.05.045
Pérez-Albaladejo, E., Fernandes, D., Lacorte, S., and Porte, C. (2017). Comparative Toxicity, Oxidative Stress and Endocrine Disruption Potential of Plasticizers in JEG-3 Human Placental Cells. Toxicol. Vitro 38, 41–48. doi:10.1016/j.tiv.2016.11.003
Peshdary, V., Hobbs, C. A., Maynor, T., Shepard, K., Gagné, R., Williams, A., et al. (2021). Transcriptomic Pathway and Benchmark Dose Analysis of Bisphenol A, Bisphenol S, Bisphenol F, and 3,3',5,5'-Tetrabromobisphenol A in H9 Human Embryonic Stem Cells. Toxicol. Vitro 72, 105097. doi:10.1016/j.tiv.2021.105097
Plastics the Facts (2019). An Analysis of European Plastics Production, Demand and Waste Data. An Analysis of European Plastics Production, Demand and Waste Data.
Powell, J. J., Faria, N., Thomas-McKay, E., and Pele, L. C. (2010). Origin and Fate of Dietary Nanoparticles and Microparticles in the Gastrointestinal Tract. J. Autoimmun. 34 (3), J226–J233. doi:10.1016/j.jaut.2009.11.006
Prata, J. C., da Costa, J. P., Lopes, I., Duarte, A. C., and Rocha-Santos, T. (2020). Environmental Exposure to Microplastics: An Overview on Possible Human Health Effects. Sci. Total Environ. 702, 134455. doi:10.1016/j.scitotenv.2019.134455
Pratiwi, A., Syafei, A. D., Assomadi, A. F., Boedisantoso, R., and Hermana, J. (2020). Microplastic Characterization Based on the Number of Occupants. AIP Conf. Proc. doi:10.1063/5.0030816
Puckowski, A., Cwięk, W., Mioduszewska, K., Stepnowski, P., and Białk-Bielińska, A. (2021). Sorption of Pharmaceuticals on the Surface of Microplastics. Chemosphere 263, 127976. doi:10.1016/j.chemosphere.2020.127976
Rafael-Vázquez, L., García-Trejo, S., Aztatzi-Aguilar, O. G., Bazán-Perkins, B., and Quintanilla-Vega, B. (2018). Exposure to Diethylhexyl Phthalate (DEHP) and Monoethylhexyl Phthalate (MEHP) Promotes the Loss of Alveolar Epithelial Phenotype of A549 Cells. Toxicol. Lett. 294, 135–144. doi:10.1016/j.toxlet.2018.05.012
Reddam, A., Mitchell, C. A., Dasgupta, S., Kirkwood, J. S., Vollaro, A., Hur, M., et al. (2019). MRNA-sequencing Identifies Liver as a Potential Target Organ for Triphenyl Phosphate in Embryonic Zebrafish. Toxicol. Sci. 172 (1), 51–62. doi:10.1093/toxsci/kfz169
Renzi, M., and Blašković, A. (2018). Litter & Microplastics Features in Table Salts from marine Origin: Italian versus Croatian Brands. Mar. Pollut. Bull. 135, 62–68. doi:10.1016/j.marpolbul.2018.06.065
Revel, M., Châtel, A., and Mouneyrac, C. (2018). Micro(nano)plastics: A Threat to Human Health. Curr. Opin. Environ. Sci. Health 1, 17–23. doi:10.1016/j.coesh.2017.10.003
Rist, S., Steensgaard, I. M., Guven, O., Nielsen, T. G., Jensen, L. H., Møller, L. F., et al. (2019). The Fate of Microplastics during Uptake and Depuration Phases in a Blue Mussel Exposure System. Environ. Toxicol. Chem. 38 (1), 99–105. doi:10.1002/etc.4285
Rochester, J. R. (2013). Bisphenol A and Human Health: A Review of the Literature. Reprod. Toxicol. 42, 132–155. doi:10.1016/j.reprotox.2013.08.008
Rogers, K. L., Carreres‐Calabuig, J. A., Gorokhova, E., and Posth, N. R. (2020). Micro‐by‐micro Interactions: How Microorganisms Influence the Fate of marine Microplastics. Limnol Oceanogr 5 (1), 18–36. doi:10.1002/lol2.10136
Santillo, D., Miller, K., and Johnston, P. (2017). Microplastics as Contaminants in Commercially Important Seafood Species. Integr. Environ. Assess. Manag. 13 (3), 516–521. doi:10.1002/ieam.1909
Sapea, S. A. f. P. b. E. A. (2019). A Scientific Perspective On Microplastics In Nature And Society. A Scientific Perspective on Microplastics in Nature and Society.
Schirinzi, G. F., Llorca, M., Seró, R., Moyano, E., Barceló, D., Abad, E., et al. (2019). Trace Analysis of Polystyrene Microplastics in Natural Waters. Chemosphere 236, 124321. doi:10.1016/j.chemosphere.2019.07.052
Schirinzi, G. F., Pedà, C., Battaglia, P., Laface, F., Galli, M., Baini, M., et al. (2020). A New Digestion Approach for the Extraction of Microplastics from Gastrointestinal Tracts (GITs) of the Common Dolphinfish (Coryphaena hippurus) from the Western Mediterranean Sea. J. Hazard. Mater. 397, 122794. doi:10.1016/j.jhazmat.2020.122794
Schirinzi, G. F., Pérez-Pomeda, I., Sanchís, J., Rossini, C., Farré, M., and Barceló, D. (2017). Cytotoxic Effects of Commonly Used Nanomaterials and Microplastics on Cerebral and Epithelial Human Cells. Environ. Res. 159, 579–587. doi:10.1016/j.envres.2017.08.043
Schymanski, D., Goldbeck, C., Humpf, H.-U., and Fürst, P. (2018). Analysis of Microplastics in Water by Micro-raman Spectroscopy: Release of Plastic Particles from Different Packaging into mineral Water. Water Res. 129, 154–162. doi:10.1016/j.watres.2017.11.011
Sedha, S., Lee, H., Singh, S., Kumar, S., Jain, S., Ahmad, A., et al. (2021). Reproductive Toxic Potential of Phthalate Compounds - State of Art Review. Pharmacol. Res. 167, 105536. doi:10.1016/j.phrs.2021.105536
Selvam, S., Manisha, A., Venkatramanan, S., Chung, S. Y., Paramasivam, C. R., and Singaraja, C. (2020). Microplastic Presence in Commercial marine Sea Salts: A Baseline Study along Tuticorin Coastal Salt pan Stations, Gulf of Mannar, South India. Mar. Pollut. Bull. 150, 110675. doi:10.1016/j.marpolbul.2019.110675
Seth, C. K., and Shriwastav, A. (2018). Contamination of Indian Sea Salts with Microplastics and a Potential Prevention Strategy. Environ. Sci. Pollut. Res. 25 (30), 30122–30131. doi:10.1007/s11356-018-3028-5
Shi, Q., Tang, J., Wang, L., Liu, R., and Giesy, J. P. (2021). Combined Cytotoxicity of Polystyrene Nanoplastics and Phthalate Esters on Human Lung Epithelial A549 Cells and its Mechanism. Ecotoxicol Environ. Saf. 213, 112041. doi:10.1016/j.ecoenv.2021.112041
Shruti, V. C., Pérez-Guevara, F., Elizalde-Martínez, I., and Kutralam-Muniasamy, G. (2020). First Study of its Kind on the Microplastic Contamination of Soft Drinks, Cold tea and Energy Drinks - Future Research and Environmental Considerations, Sci. Total Environ., 726. 138580. doi:10.1016/j.scitotenv.2020.138580
Shruti, V. C., Pérez-Guevara, F., and Kutralam-Muniasamy, G. (2020). Metro Station Free Drinking Water fountain- A Potential "microplastics Hotspot" for Human Consumption. Environ. Pollut. 261, 114227. doi:10.1016/j.envpol.2020.114227
Sicińska, P. (2018). Di-n-butyl Phthalate, Butylbenzyl Phthalate and Their Metabolites Induce Haemolysis and Eryptosis in Human Erythrocytes. Chemosphere 203, 44–53. doi:10.1016/j.chemosphere.2018.03.161
Smith, M., Love, D. C., Rochman, C. M., and Neff, R. A. (2018). Microplastics in Seafood and the Implications for Human Health. Curr. Envir Health Rpt 5 (3), 375–386. doi:10.1007/s40572-018-0206-z
Soltani, N. S., Taylor, M. P., and Wilson, S. P. (2021). Quantification and Exposure Assessment of Microplastics in Australian Indoor House Dust. Environ. Pollut. 283, 117064. doi:10.1016/j.envpol.2021.117064
Stock, V., Laurisch, C., Franke, J., Dönmez, M. H., Voss, L., Böhmert, L., et al. (2021). Uptake and Cellular Effects of PE, PP, PET and PVC Microplastic Particles. Toxicol. Vitro 70, 105021. doi:10.1016/j.tiv.2020.105021
Su, L., Deng, H., Li, B., Chen, Q., Pettigrove, V., Wu, C., et al. (2019). The Occurrence of Microplastic in Specific Organs in Commercially Caught Fishes from Coast and Estuary Area of east China. J. Hazard. Mater. 365, 716–724. doi:10.1016/j.jhazmat.2018.11.024
Teng, J., Wang, Q., Ran, W., Wu, D., Liu, Y., Sun, S., et al. (2019). Microplastic in Cultured Oysters from Different Coastal Areas of China. Sci. Total Environ. 653, 1282–1292. doi:10.1016/j.scitotenv.2018.11.057
Tickner, J. A., Schettler, T., Guidotti, T., McCally, M., and Rossi, M. (2001). Health Risks Posed by Use of Di-2-ethylhexyl Phthalate (DEHP) in PVC Medical Devices: A Critical Review. Am. J. Ind. Med. 39 (1), 100–111. doi:10.1002/1097-0274(200101)39:1<100:aid-ajim10>3.0.co;2-q
Tong, H., Jiang, Q., Hu, X., and Zhong, X. (2020). Occurrence and Identification of Microplastics in Tap Water from China. Chemosphere 252, 126493. doi:10.1016/j.chemosphere.2020.126493
Torres, F. G., Dioses-Salinas, D. C., Pizarro-Ortega, C. I., and De-la-Torre, G. E. (2021). Sorption of Chemical Contaminants on Degradable and Non-degradable Microplastics: Recent Progress and Research Trends. Sci. Total Environ. 757, 143875. doi:10.1016/j.scitotenv.2020.143875
Van Cauwenberghe, L., Claessens, M., Vandegehuchte, M. B., and Janssen, C. R. (2015). Microplastics Are Taken up by Mussels (Mytilus edulis) and Lugworms (Arenicola marina) Living in Natural Habitats. Environ. Pollut. 199, 10–17. doi:10.1016/j.envpol.2015.01.008
Van Cauwenberghe, L., and Janssen, C. R. (2014). Microplastics in Bivalves Cultured for Human Consumption. Environ. Pollut. 193, 65–70. doi:10.1016/j.envpol.2014.06.010
Vasconcelos, A. L., Silva, M. J., and Louro, H. (2019). In Vitro exposure to the Next-Generation Plasticizer Diisononyl Cyclohexane-1,2-Dicarboxylate (DINCH): Cytotoxicity and Genotoxicity Assessment in Human Cells. J. Toxicol. Environ. Health A 82 (9), 526–536. doi:10.1080/15287394.2019.1634376
Vianello, A., Jensen, R. L., Liu, L., and Vollertsen, J. (2019). Simulating Human Exposure to Indoor Airborne Microplastics Using a Breathing Thermal Manikin. Sci. Rep. 9 (1), 8670. doi:10.1038/s41598-019-45054-w
Wang, C., Chen, H., Li, H., Yu, J., Wang, X., and Liu, Y. (2020). Review of Emerging Contaminant Tris(1,3-Dichloro-2-Propyl)phosphate: Environmental Occurrence, Exposure, and Risks to Organisms and Human Health. Environ. Int. 143, 105946. doi:10.1016/j.envint.2020.105946
Wang, L., Wang, Y., Xu, M., Ma, J., Zhang, S., Liu, S., et al. (2021). Enhanced Hepatic Cytotoxicity of Chemically Transformed Polystyrene Microplastics by Simulated Gastric Fluid. J. Hazard. Mater. 410, 124536. doi:10.1016/j.jhazmat.2020.124536
Wang, X., Li, C., Liu, K., Zhu, L., Song, Z., and Li, D. (2020). Atmospheric Microplastic over the South China Sea and East Indian Ocean: Abundance, Distribution and Source. J. Hazard. Mater. 389, 121846. doi:10.1016/j.jhazmat.2019.121846
Wang, Y., Xu, L., Peng, L., Fang, C., Qin, Q., Lv, X., et al. (2021). Polybrominated Diphenyl Ethers Quinone-Induced Intracellular Protein Oxidative Damage Triggers Ubiquitin-Proteasome and Autophagy-Lysosomal System Activation in LO2 Cells. Chemosphere 275, 130034. doi:10.1016/j.chemosphere.2021.130034
Wright, S. L., and Kelly, F. J. (2017). Plastic and Human Health: A Micro Issue. Environ. Sci. Technol. 51 (12), 6634–6647. doi:10.1021/acs.est.7b00423
Wu, B., Wu, X., Liu, S., Wang, Z., and Chen, L. (2019). Size-dependent Effects of Polystyrene Microplastics on Cytotoxicity and Efflux Pump Inhibition in Human Caco-2 cells. Chemosphere 221, 333–341. doi:10.1016/j.chemosphere.2019.01.056
Wu, S., Wu, M., Tian, D., Qiu, L., and Li, T. (2020). Effects of Polystyrene Microbeads on Cytotoxicity and Transcriptomic Profiles in Human Caco‐2 Cells. Environ. Toxicol. 35 (4), 495–506. doi:10.1002/tox.22885
Xiong, H., Huang, Y., Mao, Y., Liu, C., and Wang, J. (2021). Inhibition in Growth and Cardiotoxicity of Tris (2-butoxyethyl) Phosphate through Down-Regulating Wnt Signaling Pathway in Early Developmental Stage of Zebrafish (Danio rerio), Ecotoxicol Environ. Saf., 208. 111431. doi:10.1016/j.ecoenv.2020.111431
Xu, M., Halimu, G., Zhang, Q., Song, Y., Fu, X., Li, Y., et al. (2019). Internalization and Toxicity: A Preliminary Study of Effects of Nanoplastic Particles on Human Lung Epithelial Cell, Sci. Total Environ., 694. 133794. doi:10.1016/j.scitotenv.2019.133794
Yan, Z., Jin, X., Liu, D., Hong, Y., Liao, W., Feng, C., et al. (2021). The Potential Connections of Adverse Outcome Pathways with the hazard Identifications of Typical Organophosphate Esters Based on Toxicity Mechanisms. Chemosphere 266, 128989. doi:10.1016/j.chemosphere.2020.128989
Zhang, L., Xie, Y., Zhong, S., Liu, J., Qin, Y., and Gao, P. (2021). Microplastics in Freshwater and Wild Fishes from Lijiang River in Guangxi, Southwest China. Sci. Total Environ. 755, 142428. doi:10.1016/j.scitotenv.2020.142428
Zhang, M., Li, J., Ding, H., Ding, J., Jiang, F., Ding, N. X., et al. (2020). Distribution Characteristics and Influencing Factors of Microplastics in Urban Tap Water and Water Sources in Qingdao, China. Anal. Lett. 53 (8), 1312–1327. doi:10.1080/00032719.2019.1705476
Zhang, Q., Zhao, Y., Du, F., Cai, H., Wang, G., and Shi, H. (2020). Microplastic Fallout in Different Indoor Environments. Environ. Sci. Technol. 54 (11), 6530–6539. doi:10.1021/acs.est.0c00087
Zhou, Q., Tian, C., and Luo, Y. (2017). Various Forms and Deposition Fluxes of Microplastics Identified in the Coastal Urban Atmosphere. Chin. Sci. Bull. 62, 3902–3909. doi:10.1360/n972017-00956
Zhu, L., Huang, X., Li, Z., Cao, G., Zhu, X., She, S., et al. (2021). Evaluation of Hepatotoxicity Induced by 2-ethylhexyldiphenyl Phosphate Based on Transcriptomics and its Potential Metabolism Pathway in Human Hepatocytes. J. Hazard. Mater. 413, 125281. doi:10.1016/j.jhazmat.2021.125281
Zimmermann, L., Dierkes, G., Ternes, T. A., Völker, C., and Wagner, M. (2019). Benchmarking the In Vitro Toxicity and Chemical Composition of Plastic Consumer Products. Environ. Sci. Technol. 53 (19), 11467–11477. doi:10.1021/acs.est.9b02293
Zitouni, N., Bousserrhine, N., Belbekhouche, S., Missawi, O., Alphonse, V., Boughatass, I., et al. (2020). First Report on the Presence of Small Microplastics (≤ 3 μm) in Tissue of the Commercial Fish Serranus Scriba (Linnaeus. 1758) from Tunisian Coasts and Associated Cellular Alterations. Environ. Pollut. 263, 114576. doi:10.1016/j.envpol.2020.114576
Keywords: nanoplastics, microplastics, human heath, toxicology, routes of exposure
Citation: Llorca M and Farré M (2021) Current Insights into Potential Effects of Micro-Nanoplastics on Human Health by in-vitro Tests. Front. Toxicology. 3:752140. doi: 10.3389/ftox.2021.752140
Received: 02 August 2021; Accepted: 13 September 2021;
Published: 29 September 2021.
Edited by:
Nan Sang, Shanxi University, ChinaReviewed by:
Luciana Dini, Sapienza University of Rome, ItalyLang Tran, Institute of Occupational Medicine, United Kingdom
Copyright © 2021 Llorca and Farré. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.
*Correspondence: Marta Llorca, bWxjcWFtQGNpZC5jc2ljLmVz; Marinella Farré, bWZ1cWFtQGNpZC5jc2ljLmVz