Plastics are commonly used in food and beverage packaging, pharmaceuticals, cosmetics, cleaning, and chemical products (Ganesh Kumar et al., 2020). Based on the polymeric material, the commonly used are polystyrenes, polyethylenes, polyethylene terephthalate (PET), high-density polyethylene (HDPE), low-density polyethylene (LDPE), carbon chloride polyvinyl (PVC), polypropylene, polycarbonates, and nylon (Ganesh Kumar et al., 2020). After being released into the environment, plastics are degraded and fragmented into smaller particles, increasing their toxic potential (González-Pleiter et al., 2019). Microplastics (MPs) are particles smaller than 5 mm, and nanoplastics (NPs) are particles smaller than 100 nm, although some authors define NPs as particles ranging from 1 to 1000 nm (Frias and Nash, 2019; Gigault et al., 2018).

MPs and NPs can be found in different environmental compartments. The air and water are the main routes through which MPs and NPs are released into the environment and can eventually enter the food chain, with the potential of bioaccumulation and biomagnification (Ng et al., 2018; González-Pleiter et al., 2019; Toussaint et al., 2019; Barbosa et al., 2020). In one day, a person can be exposed to 26–130 airborne MPs, representing a risk to human health, particularly in susceptible individuals (Prata, 2018). This risk is mainly associated with the potential of MPs to interact with the organic material in the respiratory system and the capacity of human alveolar epithelial cells to internalize NPs in a size-dependent manner (Prata, 2018; Xu et al., 2020).

NPs are more harmful than MPs once they can cross biological membranes easily (EFSA, 2016). Smaller NPs are more easily internalized into the cytoplasm by phagocytosis, being capable of affecting cell viability, activating the transcription of inflammatory genes, altering the expression of proteins associated with apoptosis (DR5 and caspases 3 and 9), and positively regulating pro-inflammatory cytokines (IL-8, NF- κβ, and TNF-α) (Xu et al., 2020). Besides the lungs, MPs have already been detected in urine (Pironti et al., 2022), feces (Schwabl et al., 2019), blood (Leslie et al., 2022), colorectal adenocarcinoma tissue (Cetin et al., 2023), male reproductive system (Zhao et al., 2023), and placenta (Ragusa et al., 2021).

Since the water filtration system is insufficient to remove these plastics (Eerkes-Medrano et al., 2015), the intake of food or water contaminated with MPs and NPs is one of the main routes of human exposure. Based on food intake, it is estimated that annual MP intake ranges from 39,000 to 52,000 particles. Drinking from bottled water can add 90,000 particles per year, compared to the 4000 MPs for tap water (Cox et al., 2019). Oral absorption occurs mainly in the intestine, although particles can interact with gastric cells (Banerjee and Shelver, 2021).

Although NPs accumulate size-dependent in human hepatocellular carcinoma cell line (HepG2), the cells showed internalization mechanisms dependent on clathrin-mediated endocytosis and positively regulated the expression of inflammatory genes, such as IL-6 and IL-8. (Forte et al., 2016). In human colorectal adenocarcinoma cells (Caco-2), NPs were internalized into the cytoplasm and nucleus and caused depolarization of the mitochondrial membrane (Wu et al., 2019), but no cytotoxicity, increased production of reactive oxygen species, and DNA strand breaks were observed (Cortés et al., 2020). Due to the accumulation of NPs in the liver, it is an important target for their toxicity (Ogawara et al., 1999). Oxidative stress caused by polystyrene NPs (50 nm) in mouse hepatocytes increases superoxide dismutase (SOD) activity and malondialdehyde (MDA) levels, increasing cellular DNA strand breaks (Zheng et al., 2019).

The relationship between the surface area and the volume of plastic fragments facilitates the absorption of contaminants on their surface (Rochman et al., 2013a). After entering the body, MPs and NPs can release absorbed contaminants, increasing human exposure to different compounds (Binelli et al., 2017). Several studies show that different contaminants can be adsorbed on plastic fragments and be found all over the planet, including bisphenol A, polychlorinated biphenyls, organochlorine pesticides, polycyclic aromatic hydrocarbon, and toxic metals (Mato et al., 2001; Rios et al., 2007; Frias et al., 2010; Rochman et al., 2013b; Chen et al., 2017; Barboza et al., 2020; Squadrone et al., 2022).

Bisphenol A (BPA) is an industrial product widely used in the production of plastic, mainly polycarbonate epoxy resin and other types of resin (Kang et al., 2006; WHO, 2010; Gallimberti et al., 2020). BPA is an endocrine disruptor linked to several adverse health effects, including obesity, diabetes mellitus, and cardiovascular disease (Glausiusz, 2014). Studies demonstrate that its metabolites show stronger estrogenic activity than BPA (Kovacic, 2010; Nakamura et al., 2011). In children, BPA combined with other endocrine disruptors has been associated with DNA damage (Rocha et al., 2018). Mice exposed to BPA for five days showed damage to the structure of liver mitochondria, increased hepatic MDA levels, decreased glutathione peroxidase (GPx) expression, and increased pro-inflammatory cytokines, such as IL-6 and TNF-α (Moon et al., 2012). In HepG2 cells, BPA decreased mitochondrial oxygen consumption rate, ATP production, and mitochondrial membrane potential (Moon et al., 2012).

Regulatory agencies have been re-evaluating the potential health effects of BPA exposure and adjusting imposed limits. In 2023, the European Food Safety Authority's Panel on Food Contact Materials reduced the acceptable daily intake from 0.05 mg/kg bw/day to 0.2 ng/kg bw/day (EFSA, 2023). In 2012, the Brazilian Regulatory Agency (ANVISA) banned manufacturing and importing baby bottles containing BPA (BRASIL, 2011), and in 2021, they revised the specific migration limit of BPA in plastic packaging intended for contact with food, reducing it from 0.6 mg/kg to 0.05 mg/kg of food (BRASIL, 2024). Therefore, BPA analogs are being used as an alternative; however, they have demonstrated similar toxicity (Michalowicz et al., 2015; Azevedo et al., 2019; Hercog et al., 2019).

Bisphenol S (BPS) is a BPA analog and has been the most used monomer in “BPA-free” products (Qiu et al., 2019; Glausiusz, 2014). The half-life of BPS is longer than that of BPA, making it less susceptible to biodegradation, being likely to accumulate and persist in the environment (Héliès-Toussaint et al., 2014; Qiu et al., 2019). BPS shows similar estrogenic and androgenic activities of BPA, leading to changes in proliferation, differentiation, and cell death, effects on changes in organ weight, serum lipid levels, reproductive outcomes, and enzyme expression (Rochester and Bolden, 2015; Azevedo et al., 2019).

Since BPA and BPS have similar metabolism and mechanisms of action, studies showing the effects of co-exposure of NPs and BPA or BPS are necessary to demonstrate whether NPs can aggravate disorders already reported for these compounds, which could represent a potential health risk. Therefore, when evaluating the safety of these compounds and proposing a substitute for BPA, it is necessary to consider the mechanisms of action of entire classes and not just the compound individually (Rochester and Bolden, 2015). For that reason, we evaluated whether 100 nm polystyrene NPs alone and associated with BPA or BPS can promote changes in metabolic activity, genotoxicity, and production of reactive oxygen species in human hepatocellular carcinoma (HepG2) cell line.