From microplastics to nanoplastics: environmental transformation pathways, detection challenges, and health implications

1 Introduction

Plastics are synthetic polymer-based materials widely used because of their low density, affordability, durability, flexibility, and ease of processing 2011). As plastic waste accumulates in the environment, larger materials progressively fragment through physical, chemical, and biological processes, generating particles across macro-, micro-, and nanoscale (Wang Liuwei et al., 2021; Rai et al., 2021).

Microplastics (MPs) and nanoplastics (NPs) are now widely detected across aquatic, terrestrial, atmospheric, and biological environments, including lakes, rivers, marine systems, sediments, soils, deep-sea environments, polar regions, and the atmosphere (Eriksen et al., 2013; Lechner et al., 2022; Lusher et al., 2015; Klein et al., 2015; Zubris and Richards, 2005; Jani and Wu, 2025; Peng et al., 2020; Citterich et al., 2023; Allen et al., 2019; Provencher et al., 2019). Their environmental behavior is influenced by polymer properties; for example, polyethylene (PE) is generally buoyant, whereas polystyrene (PS) is more brittle and can exhibit relatively high sorption potential (Figure 1).

Microplastics are plastic particles under 5 mm, from direct release or degradation of larger plastics. Primary MPs are intentionally small, while secondary MPs form through environmental breakdown. (Liu et al., 2024; Yang et al., 2021; Nizzetto et al., 2016; Wu et al., 2019b). They occur as fragments, pellets, filaments, and granules and are composed of polymers such as polyethylene, polypropylene, polystyrene, PVC, nylon, cellulose acetate, and PET (Hidalgo-Ruz et al., 2012; Campanale et al., 2020).

The definition of nanoplastics remains less consistent. NPs are broadly described as plastic particles below 1 μm, although studies variously use boundaries of <100 nm or up to 1000 nm (Koelmans et al., 2015; Frias and Roisin, 2019; Gigault et al., 2018). This variation reflects differences in conceptual definitions, research objectives, and analytical approaches and represents an important methodological source of heterogeneity. NPs may originate from intentionally produced nanoscale materials or from the degradation of larger plastics. Their small dimensions and high surface-area-to-volume ratio can modify mobility, environmental fate, bioavailability, and interactions with biological systems. Accordingly, this review uses the term NP operationally according to the size range reported in each cited study while recognizing the implications of differing size definitions.

The transformation of MPs into smaller particles involves more than size reduction. Weathering, fragmentation, oxidation, and biological processes change particle shape, surface chemistry, aggregation, mobility, and biological interactions (Gigault et al., 2018; Liu and Zheng, 2025). Both MPs and NPs can interact with environmental contaminants and enter food webs (Wang Yung-Li et al., 2021; Galloway et al., 2017; Ng et al., 2018). For NPs, nanoscale dimensions and reactive surfaces may facilitate interactions with biological membranes; however, observed effects vary with particle size, composition, surface chemistry, aging, exposure conditions, dose, and biological model (EFSA Panel on Contaminants in the Food Chain, 2016; Abu-Zurayk et al., 2025a). Consequently, evidence of ecological and health effects remains subject to substantial uncertainty (Rai et al., 2021).

Despite growth in plastic-pollution research, microplastics are better studied than their transformation into nanoplastics and its environmental and health implications. Changes during aging affect particle behavior, detection, exposure, and toxicity, so MP study findings cannot be directly applied to NPs.

This review examines the continuum linking environmental transformation and fate with analytical detection, exposure, toxicity, and risk assessment of MPs and NPs. A comprehensive literature search was conducted using Scopus, Web of Science, PubMed, and Google Scholar for publications from 2001 to 2026, using combinations of terms related to microplastics, nanoplastics, transformation, weathering, fragmentation, aging, environmental fate, exposure, human health, toxicity, risk assessment, detection, identification, characterization, and analytical techniques. Additional searches addressed specific environmental compartments, biological effects, and analytical methods.

The central contribution of this review is its integrated analysis of how environmental transformation influences the subsequent stages of the evidence chain. Rather than treating transformation, detection, exposure, toxicity, and risk assessment as independent topics, it examines their interconnections and identifies how changes in particle size, shape, surface chemistry, aggregation, and mobility can affect analytical recovery and quantification, exposure estimates, and interpretation of toxicological outcomes. This approach highlights how uncertainties originating during environmental transformation or analytical measurement can propagate into risk assessment. By tracing these connections, the review identifies critical evidence gaps and methodological limitations that currently constrain reliable evaluation of MP and NP risks.

2 Sources and environmental transformation pathways

Microplastics (MPs) originate from human activities and enter terrestrial, freshwater, and marine environments through direct release or the breakdown of larger plastic items. Their environmental behavior is influenced by polymer properties, including density, chemical structure, surface characteristics, and resistance to degradation. For example, polyethylene (PE) is generally buoyant, whereas polystyrene (PS) is more brittle and may exhibit relatively high sorption potential (Figure 1). Once released, MPs can undergo physical, chemical, and biological transformations that alter their size, shape, surface chemistry, and environmental behavior.

2.1 Sources of microplastics

Microplastics (MPs) originate from human activities and enter environments via direct release or breakdown of larger plastics. They are either primary, released directly as small particles, or secondary, formed from larger item fragmentation. MPs can change in size, shape, surface chemistry, and behavior due to physical, chemical, and biological processes. Figure 2 illustrates primary sources and how larger plastics transform into secondary fragments.

2.1.1 Primary sources

Primary microplastics are intentionally produced or released as small particles during the use or processing of plastic products. Examples include synthetic textile fibers from laundering and particles from industrial and domestic activities. Wastewater is a major pathway for these particles, carrying them into aquatic environments through municipal, industrial, and laundry effluents. (Monira et al., 2023a; Qian et al., 2024b; Wilkes et al., 2024b).

Wastewater contributes to microplastic transport since treatment often cannot fully remove tiny particles. MPs and NPs can thus pass through systems and enter waters, where they may change and spread. (Murray and Örmeci, 2020; Kirstein et al., 2016a; Eerkes-Medrano et al., 2015; Marfella et al., 2024).

2.1.2 Secondary sources

Secondary microplastics form from the fragmentation and degradation of larger plastic materials once released into the environment. Key sources include plastic packaging, agricultural and fishing materials, and gardening plastics. Environmental stressors like UV radiation, abrasion, temperature changes, moisture, chemical reactions, and biological activity gradually weaken these materials, leading to the creation of smaller plastic particles.

Agricultural plastics represent an important terrestrial source because materials such as mulch films and other plastic products can remain in soil and undergo long-term environmental weathering. Their degradation may facilitate the formation and accumulation of micro-and nanoscale plastic particles in agricultural environments (Alvarado-Zambrano et al., 2023).

Similarly, fishing and aquaculture activities can contribute to the accumulation of plastic debris in coastal environments. Weathering and mechanical degradation of these materials can subsequently contribute to the formation of secondary microplastics (Zhang et al., 2025).

Long-term degradation of gardening and landscape plastics provides a further example of secondary microplastic formation (Sobhani et al., 2021). revealed that common gardening materials like bubble wrap and landscape fabrics can weather and abrasively break down into particles from millimeters to submicrons. This suggests that larger plastic materials can lead to microplastic formation and even produce nanoscale particles.

The distinction between primary and secondary sources is crucial for understanding the environmental presence of microplastics (MPs) and nanoplastics (NPs). Primary particles are smaller at entry, while secondary particles form through the degradation of larger plastics. Table 1 summarizes the sources identified in the review, with further discussion on the transformation processes that lead to size reduction in the following subsections.

TABLE 1. Sources of microplastics and nanoplastics identified in the review.

2.2 Physical transformation pathways

The formation of microplastics and nanoplastics occurs through the transfer of environmental energy to polymers, resulting in mechanical stress and bond weakening. Key drivers include ultraviolet (UV) radiation, mechanical abrasion, and thermal fluctuations, which cause surface embrittlement and fragmentation into smaller particles.

2.2.1 Ultraviolet (UV) radiation and surface photo-oxidation

UV radiation accelerates plastic weathering by inducing photo-oxidation, creating free radicals and oxygen groups like carbonyls and hydroxyls. This causes chain scission, increasing surface polarity and decreasing mechanical strength, making plastics prone to microcracks and breaking. (Gao et al., 2025). UV-induced degradation varies with polymer, additives, and irradiation. Different UV sources like UVA lamps and xenon-arc systems cause distinct degradation pathways due to spectral differences. Stabilizers like antioxidants delay oxidation, resulting in non-linear aging indicators such as the carbonyl index. Interpreting photodegradation requires considering material formulation and environment. (Gao et al., 2025).

2.2.2 Mechanical abrasion as a fragmentation amplifier

Mechanical abrasion is crucial in forming micro- and nanoplastics in nature. External forces such as wave action and sediment impacts stress on polymer surfaces, creating scratches, pits, and grooves that lead to cracking and fragmentation. Research indicates UV-weathered plastics have lower resistance than unaged plastics, showing how photochemical degradation and mechanical stress speed up fragmentation (Rostampour et al., 2025; Burrows et al., 2024).

2.2.3 Thermal and moisture-driven modulation of physical weathering

Thermal conditions and environmental moisture are crucial in physical weathering, influencing processes together. Higher temperatures boost polymer mobility and speed up photochemical and hydrolytic reactions, while moisture impacts surface swelling, hydrolysis, and crack propagation. Research indicates that combined UV radiation, high temperatures, and humidity lead to increased surface cracking and microplastic formation, particularly in PET. Elevated temperatures promote particle aggregation, while high humidity increases surface roughness and crack density. Therefore, thermal and moisture conditions govern degradation rates and affect particle morphology and size distribution (Rostampour et al., 2025). Table 2 clarifies the Physical transformation pathways of microplastics and nanoplastics and their environmental effects.

TABLE 2. Physical transformation pathways of microplastics and nanoplastics and their environmental effects.

2.2.4 Integration of physical pathways in environmental contexts

In natural environments, physical weathering processes are interconnected. UV radiation causes chemical degradation and surface weakening, making materials prone to mechanical breakdown. Temperature and moisture influence crack formation, polymer mobility, and particle detachment. Over time, plastics degrade into micro- and nanoplastics, increasing their surface area and reactivity, affecting chemical aging, contaminant adsorption, and biological interactions. Understanding these processes is vital for modeling plastic fate and ecological risks (Rostampour et al., 2025; Burrows et al., 2024).

2.3 Chemical transformation pathways: oxidation and hydrolysis

Plastic degradation involves chemical processes like oxidation and hydrolysis, causing polymer chain scission and material changes. These chemical shifts weaken the polymer, leading to mechanical breakdown (Pfohl et al., 2025; Shi et al., 2022). As a result, surface polarity, hydrophilicity, and charge distribution are altered, which subsequently influences particle behavior and environmental interactions (Lu et al., 2023). Hydrolysis is a crucial degradation process in moisture-rich environments such as soils and wastewater, where water molecules break bonds in hydrolysable polymers, weakening structures alongside oxidation. Chemical degradation arises from multiple interacting pathways rather than isolated reactions. (Pfohl et al., 2025).

Environmental aging alters surface chemistry, affecting adsorption behavior and reactivity towards contaminants. These changes are well-documented in aged plastics versus pristine materials, highlighting the link between chemical weathering and surface processes (Lu et al., 2023). Chemical transformation processes are affected by environmental factors like UV intensity, oxygen availability, temperature, and moisture, along with polymer-specific properties such as crystallinity and chemical structure. Together, these elements influence the rate and extent of degradation in various environmental compartments (Atugoda et al., 2023; Gigault et al., 2018). Table 3 demonstrates the Chemical transformation pathways of micro- and nanoplastics and their environmental consequences.

TABLE 3. Chemical transformation pathwaysof micro- and nanoplastics and their environmental consequences.

2.4 Biological transformation pathways: microbial degradation and enzymatic activity

Biological transformation involves interactions between micro- and nanoplastics and microorganisms or enzymes that can alter polymer chains through surface erosion, chain scission, and partial mineralization. Degradation depends on enzyme specificity, microbial colonization, biofilm formation, and environmental conditions, which collectively influence transformation rates and the persistence of plastics.

2.4.1 Microbial colonization and biofilm-assisted degradation

Microorganisms initially attach to polymer surfaces and form biofilms that concentrate cells and enzymes at the polymer–water interface, increasing enzymatic activity (Choi et al., 2024). Extracellular polymeric substances (EPSs) alter surface properties and enable localized chemical disruption, especially in polymers with hydrolyzable functional groups like polyethylene terephthalate (PET). Enzymatic hydrolysis of PET can yield low-molecular-weight products such as terephthalic acid and ethylene glycol (Yousafzai et al., 2025). Figure 3 illustrates the complete process of plastics biodegradation.

2.4.2 Enzymatic mechanisms and polymer-specific reactivity

Microbial degradation mainly occurs through enzymatic hydrolysis and sometimes oxidation. PETase and MHETase hydrolyze PET ester bonds to produce intermediates like MHET, which undergo further metabolism (Choi et al., 2024). Enzyme adsorption and efficiency depend on polymer hydrophobicity, crystallinity, and surface roughness (Sathish et al., 2025). Abiotic weathering, such as UV and abrasion, increases surface roughness and heterogeneity, aiding microbial attachment and enzymatic access. Abiotic and biological processes often act together in degradation. Table 4 summarizes key biological pathways.

TABLE 4. Biological transformation pathways.

2.4.3 Environmental microbiomes and degradation efficiency

Biodegradation efficiency varies substantially with microbial community composition, environmental conditions, and polymer chemistry. In estuarine sediments and landfill leachates, diverse microbial communities degraded biodegradable polycaprolactone (PCL) by >90% within 50–60 days under aerobic and thermophilic conditions, whereas resistant polymers such as polyethylene (PE) and PET showed minimal or no degradation (Pires et al., 2024). These findings highlight the importance of polymer chemistry and environmental conditions, particularly temperature regimes, in determining biodegradation outcomes.

3 Environmental distribution and fate

Figure 4 illustrates the lifecycle of plastics from release as primary or secondary particles through environmental transport, transformation, biodegradation, and potential entry into food webs. The environmental distribution and fate of MPs and NPs are governed by coupled transport, transformation, and partitioning processes. These processes depend on particle size, density, polymer composition, surface chemistry, and environmental conditions including salinity, ionic strength, organic matter, and biological activity.

3.1 Environmental transport and redistribution

Following release, MPs and NPs can be redistributed among terrestrial, atmospheric, and aquatic compartments through runoff, rivers, wastewater, atmospheric circulation, and sedimentation. Land-based activities are major sources, while rivers provide important pathways for transferring plastic debris from terrestrial systems to marine environments (Li et al., 2016; Lebreton et al., 2018). Transport is therefore not unidirectional; particles can be deposited, remobilized, and transferred repeatedly between environmental compartments.

Particle characteristics influence transport behavior. Larger and denser particles are more readily affected by gravitational settling, whereas smaller particles, particularly NPs, are more strongly influenced by Brownian motion and colloidal interactions. Atmospheric transport can further facilitate long-range redistribution, with wind-driven resuspension and subsequent precipitation allowing plastic particles to reach areas far from their sources (Zhang Yulan et al., 2020; Padha et al., 2022).

3.2 Physicochemical controls on mobility and partitioning

The mobility of MPs and NPs is controlled by interactions between particle properties and environmental conditions. Aging, mineral association, organic matter, and contaminant adsorption can modify particle surfaces and density, thereby changing their transport and deposition behavior (Chen et al., 2023; Alimi et al., 2018).

For NPs, aggregation is particularly important. Interactions with dissolved and particulate organic matter can promote homo- and heteroaggregation, increasing effective particle size and enhancing sedimentation (Shiu et al., 2020). Water chemistry regulates these interactions; increased salinity and ionic strength reduce electrostatic repulsion, while multivalent metal cations can further destabilize colloidal suspensions and promote aggregation (Cui et al., 2018; Sharma et al., 2021; García-Rollán et al., 2025).

Surface chemistry also determines interactions with environmental contaminants. The high surface-area-to-volume ratio of NPs provides extensive adsorption sites, while weathering can introduce oxygen-containing functional groups and alter surface reactivity. Hydrophobic polymer surfaces can additionally interact with hydrophobic organic contaminants, allowing plastic particles to influence contaminant transport and partitioning (Mato et al., 2001).

3.3 Transformation and biological modification

Transport occurs alongside continuous particle transformation. Sunlight, temperature fluctuations, oxidation, and mechanical forces can weaken polymer structures and promote fragmentation, potentially generating secondary micro- and nanoplastics (Liu and Zheng, 2025). These changes can increase surface area and modify surface chemistry, thereby altering subsequent mobility and contaminant interactions.

Biological processes provide an additional transformation pathway. Microbial colonization produces biofilms, commonly termed the “plastisphere,” which modify particle hydrophobicity, surface roughness, effective density, and reactivity (Jacquin et al., 2019; Richard et al., 2019; Gong et al., 2019). Biofouling can increase the density of initially buoyant particles and promote their transfer toward deeper waters and sediments (Amaral-Zettler et al., 2021). The plastisphere may also provide habitats for diverse microorganisms, potentially influencing microbial dispersal associated with plastic particles (Lafferty et al., 2015; Molenaar et al., 2021).

Consequently, environmental fate is dynamic rather than determined by a single transport pathway. Weathering can alter particle size and surface properties; aggregation and biofouling can modify effective size and density; and deposition can transfer particles into soils and sediments, where subsequent remobilization may return them to the water column or atmosphere. These coupled processes determine the persistence, distribution, and exposure potential of MPs and NPs.

4 Case studies of “hotspot” environments

Because transport and transformation occur across connected environmental compartments, some locations function as important hotspots where MPs and NPs accumulate, transform, or are released to downstream environments. The following case studies illustrate how these general processes manifest in wastewater, soils, sediments, aquifers, and remote environments.

4.1 Wastewater

Wastewater treatment systems receive plastic particles from municipal, industrial, and laundry sources and therefore represent important interception points between terrestrial sources and receiving waters. Although treatment can remove a substantial fraction of plastic particles, smaller particles may remain in treated effluents or become concentrated in sludge (Wilkes et al., 2024a; Monira et al., 2023b; Qian et al., 2024a; Marfella et al., 2024).

A Saudi Arabian sewage-treatment study investigated nanoplastics ranging from 50 to 2500 nm using nano-flow cytometry. Particles <100 nm accounted for 44% of detected particles, while the proportion of plastic particles increased from 16% in raw sewage to 41% in effluent. Micro-Raman spectroscopy and SEM-EDX identified polystyrene, PVC, polyethylene, PTFE, polyamide, and polypropylene. The increase in the relative proportion of plastic particles in effluent suggests that treatment does not uniformly remove particles and may preferentially retain larger or denser fractions. The persistence of plastics in treated wastewater is particularly relevant to water reuse in arid regions (Iskandarani et al., 2025).

4.2 Soils

Soils receive MPs and NPs through agricultural plastics, irrigation, wastewater, sewage sludge, and other land-use activities. Their retention can create long-term reservoirs, while agricultural activity, runoff, and erosion provide opportunities for redistribution (Uddin et al., 2026a; Pérez-Reverón et al., 2023).

Field evidence demonstrates that plastic products can generate secondary terrestrial sources. In a 7-year study, weathering of bubble wrap and landscape fabrics produced microplastics and near-nanoscale particles in garden soils, confirming that commonly used gardening materials can contribute to plastic accumulation in terrestrial environments (Sobhani et al., 2021).

4.3 Sediments

Sediments represent important accumulation zones because particles transported through aquatic systems can be retained following deposition. This creates potential exposure for benthic organisms and provides a reservoir from which particles may subsequently be remobilized.

Experimental evidence demonstrates that sediment characteristics influence this retention. Polystyrene nanoplastics were more mobile in freshwater than marine sediments, where higher salinity and finer grains promoted particle retention. Humic acid had little influence on mobility in freshwater but enhanced transport in marine sediments, indicating that interactions among salinity, organic matter, and particle–sediment interactions can produce contrasting transport behavior (Sun et al., 2024).

Coastal sediments may receive particularly high plastic loads. In Laizhou Bay, microplastic concentrations ranged from 36 to 2125.5 items/g dry weight, averaging 534.7 items/g. Fragments (61.91%) and particles (37.45%) predominated, with polyamide (32.73%), PET (12.42%), and PU (10.81%) as the major polymers. Aquaculture and fishing were identified as important sources. Although the pollution load index indicated relatively low overall ecological risk, the polymer risk index highlighted concerns associated with PU and PVC (Zhang et al., 2025).

4.4 Aquifers

Aquifers are emerging hotspots because groundwater can receive plastic particles through agricultural leaching, septic systems, and surface infiltration. Their relative isolation from direct sunlight and slower transformation rates may allow particles to persist once transported into subsurface environments.

A coastal aquifer in northwest Mexico contained 10–34 microplastic particles L^-1, dominated by iPP, HEC, PVC, and LDPE. Agriculture and septic outflows were identified as major sources, with soil infiltration and marine intrusion proposed as transport pathways (Alvarado-Zambrano et al., 2023). Although the study focused on microplastics, continued degradation of retained particles could provide a potential pathway for in situ nanoplastic formation (Lavers and Bond, 2017; Chan et al., 2011; Moeck et al., 2023). These findings indicate that groundwater systems warrant greater attention in assessments of subsurface plastic contamination.

4.5 Atmospheric and remote environments

Remote environments provide evidence that plastic pollution can extend well beyond areas of direct human activity. Atmospheric deposition can transfer particles to high-altitude and cryospheric environments, where they may be temporarily retained before subsequent redistribution.

Glaciers, for example, can act as reservoirs of atmospheric plastic particles. Melting may release accumulated particles into downstream aquatic systems, linking atmospheric, cryospheric, and aquatic compartments (Belioka and Achilias, 2024). Atmospheric exposure may also modify plastic particles through photochemical processes, potentially contributing to fragmentation during transport (Liu and Zheng, 2025).

These observations demonstrate that environmental hotspots are not isolated reservoirs. Rather, they represent temporary accumulation points within a connected system in which particles can be transported, transformed, deposited, and subsequently remobilized.

5 Exposure pathways and metrics

Micro- and nanoplastics are pervasive in air, water, and soil, affecting living organisms through interactions with biotic surfaces like the plastisphere. Understanding exposure pathways is essential, but assessing this exposure is difficult due to inconsistent methods, limited environmental data, and unclear regulations on plastic production and lifecycle management (Lane et al., 2025). Figure 5 illustrates the complex transport and transformation of micro- and nanoplastics in various environments, highlighting key mechanisms such as aggregation and co-contaminant sorption that affect their bioavailability and exposure in ecological and human systems.

5.1 Ecosystems: plants, soil organisms, aquatic food webs

MPs and NPs are widespread in air, water, and soil and can enter organisms through direct exposure and transfer across food webs. Exposure assessment remains challenging because of heterogeneous sampling and analytical methods, limited environmental data, and inconsistent reporting practices (Lane et al., 2025). Figure 5 illustrates major environmental transport and transformation processes, including aggregation and co-contaminant interactions, that influence plastic bioavailability and exposure.

MPs and NPs enter terrestrial ecosystems through atmospheric deposition, sewage sludge, agricultural practices, wastewater, plastic mulching, landfill emissions, and runoff. In soils, particles can be ingested by soil organisms such as earthworms, facilitating redistribution within the soil matrix (Adhikari, 2024; Rajput et al., 2026). Plants may also take up plastic particles through roots via crack-entry, endocytosis, and apoplastic or symplastic transport, followed by translocation to stems, leaves, and potentially edible tissues (Yu et al., 2024a).

Wastewater treatment plants (WWTPs) represent an important interface between anthropogenic sources and natural ecosystems. Although conventional treatment can remove >90% of plastic particles in many plants, removal is incomplete and substantial quantities may remain in effluent or accumulate in sludge (Komorowska-Kaufman and Marciniak, 2024; Reza et al., 2024). Sludge application to agricultural soils can subsequently reintroduce MPs and NPs into terrestrial systems and facilitate their transfer through runoff (Bodzek et al., 2024).

Agricultural plastic mulching is another important source. Residual PE, PP, and PVC films left after harvesting can fragment into MPs and NPs, which may persist in soil and influence soil aggregation, porosity, water retention, nutrient cycling, and plant growth (Steinmetz et al., 2016; Qi et al., 2018; Salama and Geyer, 2023; Liu et al., 2023a; Machado et al., 2019). Evidence that NPs can enter plant tissues and move through vascular systems raises concerns about their transfer into terrestrial food webs (Gao et al., 2023; Sembada and Wuled Lenggoro, 2024; Su et al., 2019; Schwab et al., 2016).

Landfills provide additional sources through plastic degradation and leachate. Rainwater percolating through waste can mobilize MPs and NPs into groundwater and surface waters, while wind can redistribute particles from exposed waste (He et al., 2019; Wojnowska-Baryła et al., 2022). Atmospheric precipitation provides another pathway, depositing plastic particles transported through the atmosphere onto terrestrial and aquatic surfaces. MPs have been detected in rain and snow, whereas NP detection remains considerably more challenging (Evangeliou et al., 2020; Allen et al., 2019; Zhang Yulan et al., 2020).

Ecosystem exposure is assessed by measuring plastics in water, sediment, soil, plants, and organism tissues and by examining their transfer through food webs. Analytical approaches include FTIR and Raman spectroscopy, electron microscopy, pyrolysis–GC/MS, nanoparticle tracking analysis (NTA), and dynamic light scattering (DLS) (Roberto de BruinCarlo et al., 2022; Huber et al., 2023). However, no single metric adequately represents exposure across the full MP–NP size range. Particle number is particularly relevant for NPs because their mass can be extremely low despite high particle abundance, whereas mass or volume may better represent accumulation and environmental loading (Thornton et al., 2022). Size distribution is also important because biological interactions and transport can vary substantially across the nanoscale range (Molenaar et al., 2021).

Accordingly, exposure characterization should consider multiple complementary metrics, including particle number, mass, size distribution, polymer composition, morphology, and, where feasible, surface area and volume (Thornton et al., 2022; Fraissinet et al., 2024). Particle aging and morphology further influence exposure by altering surface reactivity, contaminant sorption, aggregation, and uptake. Table 5 summarizes representative ecosystem exposure pathways.

TABLE 5. Relevant examples of exposure pathways in ecosystems to the plastic particles.

5.2 Humans: ingestion, inhalation, and dermal exposure

Human exposure to MPs and NPs occurs primarily through ingestion, inhalation, and, to a lesser extent, dermal contact (Cox et al., 2019; Prata et al., 2020). Plastic particles have been reported in human-associated samples and biological compartments, including feces, blood, lung tissue, breast milk, and placenta, demonstrating widespread exposure (Leslie et al., 2022; Ragusa et al., 2021; Fujimoto et al., 2022). However, detection does not by itself establish biological harm, and the extent to which detected particles represent internal exposure remains an important research question.

Ingestion occurs through contaminated food and drinking water. MPs and NPs have been reported in seafood, fruits, vegetables, honey, bottled water, and other food and beverage sources. Atmospheric deposition can additionally contaminate food and water, while agricultural soils provide a potential pathway for particle transfer into crops and subsequently into food webs (Cox et al., 2019; Jasso–Salcedo et al., 2024; Smith et al., 2018).

Inhalation is another important exposure route. MPs and NPs can become airborne through tire wear, textile fibers, degraded plastics, building materials, and household dust. Indoor environments may represent important exposure settings because of synthetic textiles, carpets, furniture, and accumulated dust (Dris et al., 2016; Prata, 2018).

Dermal exposure is generally considered less important than ingestion and inhalation because intact skin provides an effective barrier against larger particles. Nevertheless, exposure can occur through contact with contaminated environments, cosmetics, personal care products, and synthetic textiles. NPs may have greater potential for skin penetration because of their small size, although the extent and biological significance of dermal uptake remain uncertain (Abdallah et al., 2015; Ageel et al., 2022; H. Leslie and Depledge, 2020).

Human exposure assessment involves sampling, particle isolation, identification, and quantification in environmental or biological matrices (Roberto de BruinCarlo et al., 2022; Zhao et al., 2025). Depending on the matrix and particle size, methods include FTIR and Raman spectroscopy, optical and electron microscopy, pyrolysis–GC/MS, and thermal analytical techniques (Rauert et al., 2025; Brits et al., 2024). Reported exposure may be expressed as particle number, mass concentration, or polymer mass fraction. For example, fecal samples may be reported as particles per Gram, while blood or water may be expressed as particles per milliliter or mass per volume.

For NPs, particle number and size distribution are particularly important because nanoscale particles can have very low mass but high abundance and biological reactivity (Dhada et al., 2023). However, inconsistent reporting units and the limited sensitivity of conventional analytical methods complicate comparisons among studies and risk assessment (Xu et al., 2026). Techniques such as NTA, TEM, DLS, and AFM may improve nanoscale characterization, but their applicability depends on particle concentration, matrix composition, size range, and the need for chemical identification (Xu et al., 2026; Christodoulou et al., 2026; Medley et al., 2023).

Thus, exposure assessment should distinguish between environmental occurrence, measured exposure, and internal biological exposure. The smaller size and distinct physicochemical behavior of NPs require particular attention to particle number, size distribution, and analytical detection limits. Figure 6 summarizes the principal human exposure pathways, while Table 6 provides representative examples.

TABLE 6. Relevant examples of human exposure pathways to the NPs and MPs.

5.3 Mismatch between laboratory exposure levels and environmentally relevant concentrations

Environmental risk assessment assumes that laboratory findings can be extrapolated to natural systems, yet this assumption is increasingly questioned because experimental particle types, concentrations, exposure durations, and conditions often differ from those in the environment (Elert et al., 2017; Lenz et al., 2016).

Laboratory studies commonly use microplastic concentrations 10²–10⁷ times higher than environmental levels to produce measurable effects within practical experimental periods, but such exposures may not represent the lower, long-term exposure experienced by organisms in nature (Zhang et al., 2026; Ivleva, 2021). Experimental particles are also often uniform materials such as pristine polystyrene microspheres, whereas environmental plastics vary in size, shape, polymer composition, and chemical properties. Weathering through oxidation, UV degradation, and biofilm formation further alters particle characteristics and reactivity (Zhang et al., 2026; Hartmann et al., 2019; Burns and Boxall, 2018).

Exposure conditions also differ substantially. Organisms in natural systems may experience chronic, low-dose exposure through multiple pathways, including ingestion with food and interactions with organic matter and sediments, whereas laboratory studies commonly use direct exposure to isolated particles over shorter periods (Zhang et al., 2026; Burns and Boxall, 2018; Koelmans et al., 2017). Environmental plastics also occur as complex mixtures with heavy metals, persistent organic pollutants (POPs), and microbial biofilms, conditions that are rarely reproduced experimentally (Rochman et al., 2013).

A major additional uncertainty is the lack of reliable environmental concentration data for nanoplastics. Their small size and current analytical limitations make detection and quantification difficult (Lenz et al., 2016; Gigault et al., 2018; Koelmans et al., 2015). Consequently, uncertainties in environmental concentrations, together with the use of pristine particles and simplified exposure systems, complicate toxicological interpretation, exposure assessment, human-health evaluation, and risk characterization. Experimental designs that better reflect environmentally relevant particle characteristics, concentrations, exposure durations, and exposure pathways are therefore needed to improve extrapolation to ecological risk assessment and support evidence-based regulation.

6 Nanoplastics analytical detection and quantification challenges

Analytical limitations cause lab-environment mismatch discussed in Section 5.2. Current detection methods struggle to quantify environmental nanoplastics and characterize weathered particles, leading to uncertainties affecting exposure, toxicity, and risk assessments.

6.1 Sample preparation

Environmental samples often contain heteroaggregates comprising NPs, other plastic particles, and organic and inorganic contaminants, making accurate NP isolation and identification challenging and increasing the risk of false-positive or false-negative results (Koelmans et al., 2019; Huang et al., 2021; Chia et al., 2022). Figure 7 summarizes the main sample-preparation steps used before NP analysis.

Quality control requires clean equipment, protective clothing, contamination-free conditions, covered samples, filtered reagents, and procedural blanks to identify contamination sources (Cai et al., 2021a; Huang et al., 2021; Ośko et al., 2025). Recovery tests and positive controls should also be incorporated into method (Prata et al., 2024; Calero et al., 2024; Nguyen et al., 2019).

6.1.1 Pretreatment

Digestion is commonly required to remove organic matter from complex matrices, although relatively clean samples such as drinking water may require limited (Ośko et al., 2025; Pérez-Reverón et al., 2023). Enzymatic digestion (Rauert et al., 2025) and chemical approaches, including acids/bases, sodium dodecyl sulfate, H₂O₂, and Fenton’s reagent, have been applied (Lin et al., 2020; Rist et al., 2017; Li et al., 2024; Brits et al., 2024; Löder et al., 2017; Okoffo and Thomas, 2024).

Alternatively, NPs can be extracted and enriched using hydrophobic functionalized magnetic nanoparticles, which capture NPs through hydrophobic interactions and allow magnetic separation (Pasanen et al., 2024; Surette et al., 2023; Wang et al., 2024). Metal–organic frameworks and layered double hydroxides have also been investigated as adsorbents for NP extraction from water (Modak et al., 2023; Liu Peng et al., 2024; Nene et al., 2025).

6.1.2 Preconcentration

Ultrafiltration (UF) reduces sample volume and enriches NPs using nanoporous membranes, typically with a 10–100 kDa molecular-weight cutoff (Li et al., 2020; Halle et al., 2017; Abu-Zurayk et al., 2025b). Cross-flow UF improves throughput and reduces pore clogging, enabling enrichment of trace NPs from large environmental sample volumes while minimizing aggregation and sample loss (Schwaferts et al., 2019; Mintenig et al., 2018).

Ultracentrifugation can efficiently sediment NPs but may promote aggregation or particle damage and lacks selectivity because co-contaminants can also be pelleted (Abdolahpur Monikh et al., 2019; Cerasa et al., 2021; Schwaferts et al., 2019). Continuous-flow centrifugation addresses the need for higher-throughput enrichment and has been applied to river-water sampling at several hundred L h^-1 (Ran et al., 2000; Hildebrandt et al., 2019). Cloud-point extraction provides another enrichment approach in which nonionic surfactants capture NPs through temperature-induced micelle formation (Xie et al., 2010).

6.1.3 Separation

6.1.3.1 Membrane filtration

Membrane filtration is simple and widely accessible for NP fractionation. For example, 0.2-µm PTFE membranes have been used to separate MPs from NPs in snow samples. However, polymeric membranes may introduce contamination, while smaller pore sizes reduce flow rates and increase fouling and pore clogging (Li et al., 2020; Abu-Zurayk et al., 2025b). Sequential filtration with decreasing pore sizes can reduce clogging (Nguyen et al., 2019), but conventional filtration remains more suitable for relatively clean laboratory samples than complex waters. Cross-flow UF combined with techniques such as asymmetric flow field-flow fractionation (AF4) can provide improved enrichment and fractionation for environmental samples (Nene et al., 2025; Schwaferts et al., 2019; Mintenig et al., 2018).

6.1.3.2 Field-flow fractionation (FFF)

Field-flow fractionation separates NPs according to size, density, or molar mass and provides high-resolution fractionation over approximately 1 nm–100 µm. Its versatility allows coupling with UV-visible, refractive-index, DLS, MALS, fluorescence, and MS detectors for simultaneous particle characterization and quantification (Cai et al., 2021a; Schwaferts et al., 2019; Wahl et al., 2021; Hadri et al., 2020; Magrì et al., 2018; Ma et al., 2010; Abdolahpur Monikh et al., 2019; Engel et al., 2014; Valido et al., 2023; Correia and Loeschner, 2018; Barber et al., 2020; Ruhland et al., 2019; Ośko et al., 2025). An integrated AF4-MALS/Py-GC-MS workflow has demonstrated the ability to fractionate NPs by size while identifying and quantifying polymers in individual fractions (Mintenig et al., 2018).

6.1.3.3 Chromatography

Size-exclusion chromatography (SEC) and hydrodynamic chromatography (HDC) provide size-based separation and can be coupled with online UV or DLS detection (Gagné, 2022; Pirok et al., 2017; Brewer and Striegel, 2010; Philippe et al., 2014). SEC generally covers approximately 1–100 nm but may cause stress-induced particle degradation, whereas HDC accommodates a broader range of approximately 1–1000 nm and can provide faster and more consistent polymer analysis (Schwaferts et al., 2019; Fu et al., 2020).

6.1.3.4 Capillary electrophoresis

Capillary electrophoresis (CE) provides rapid, cost-effective, high-resolution separation of polydisperse nanoparticle samples based primarily on surface charge and size. It has potential for characterizing both primary and environmentally transformed NPs, including oxidized particles with altered surface properties (Adam and Vaculovicova, 2017; Schwaferts et al., 2019). UV-visible, laser-induced fluorescence, and MS detectors can be coupled with CE for NP identification and quantification (Adam and Vaculovicova, 2017; Ośko et al., 2025).

Sample preparation varies with matrix complexity, polydispersity, and concentration. Some workflows need minimal treatment, but multi-step procedures may cause analyte loss and false negatives (Ośko et al., 2025; Nguyen et al., 2019). Therefore, the matrix determines the preparation strategy, which should minimize steps and maximize polymer recovery.

Beyond matrix-related challenges, the operational definition of NPs introduces another important source of methodological variability that can affect their detection, quantification, and interpretation.

6.2 Methodological implications of nanoplastic size definitions

Variations in the size ranges used to define nanoplastics (NPs) create significant methodological inconsistencies. These size boundaries affect which particles are detectable and quantifiable, impacting the concentrations reported in environmental samples. Additionally, toxicological studies that use different NP sizes may examine particles with varying surface areas, aggregation behaviors, and bioavailability, leading to differing biological effects. Such discrepancies complicate comparisons across studies, as differences in concentrations or responses may stem from size range variations and detection limits rather than actual differences in environmental presence or toxicity. Thus, it is crucial to clearly report and consider the operational size range when interpreting analytical, environmental, and toxicological results.

6.3 Nanoplastics detection, identification, and quantification

Nanoplastic analysis entails assessing physical properties like particle size, distribution, concentration, and morphology, alongside chemical characteristics such as polymer identity and mass concentration. Because no single instrument can measure everything at the nanoscale, complementary analytical techniques are essential. Additionally, detecting nanoscale particles does not confirm their identity as nanoplastics (Mariano et al., 2021).

6.3.1 Polymer identification and chemical characterization

6.3.1.1 FTIR

Fourier-transform infrared (FTIR) spectroscopy is widely used for polymer identification because of its chemical specificity, but its application to nanoplastics is constrained by the infrared diffraction limit. Its spatial resolution is typically about 10–20 μm, making reliable identification of particles below 1 μm generally difficult (Xie et al., 2024). AFM–IR can overcome this limitation by combining nanoscale morphological imaging with infrared spectra, enabling characterization of nanoplastics as small as 20 nm (Nene et al., 2025).

6.3.1.2 Raman

Raman spectroscopy provides non-destructive polymer identification through molecular fingerprints and has a spatial resolution of approximately 1 μm. However, the weak Raman scattering of nanoplastics limits conventional Raman sensitivity, prompting the use of surface-enhanced Raman spectroscopy (SERS) to improve detection (Mogha and Shin, 2023). Coupling Raman with SEM has enabled characterization of PVC nanoplastics as small as 360 nm (Zhang Wen et al., 2020). Remaining challenges include fluorescence interference, limited sensitivity, spatial-resolution constraints, and the need for reproducible plasmonic substrates in SERS applications (Sobhani et al., 2020; Luo et al., 2023; Wetzel, 2026; Mogha and Shin, 2023).

6.3.1.3 Py-GC/MS

Pyrolysis–gas chromatography–mass spectrometry (Py-GC/MS) identifies and quantifies plastics by thermally decomposing polymers into volatile products, which are separated chromatographically and identified by mass spectra (Kusch, 2017; Li Jingyi et al., 2018; Li Jia et al., 2018). It requires limited pretreatment and provides high sensitivity, with polymer masses detectable at nanogram levels in complex matrices (Kusch, 2017; Gahn et al., 2026).

However, thermal degradation can produce overlapping products, while natural organic matter may generate interfering pyrolysis compounds, requiring careful marker selection and sample purification. The relatively small sample mass analyzed (∼0.5 mg) can also limit application to dilute aqueous samples (Cai et al., 2021a). As a destructive, mass-based technique, Py-GC/MS provides limited information on particle number, size, or morphology.

6.3.2 Particle detection, morphology, and size characterization

6.3.2.1 Electron microscopy

Scanning electron microscopy (SEM) and transmission electron microscopy (TEM) provide high-resolution visualization of nanoplastic size, morphology, and surface features. SEM can be coupled with energy-dispersive X-ray spectroscopy (EDS) for elemental analysis, while TEM provides higher-resolution imaging and can be combined with EDS or electron energy-loss spectroscopy (EELS) for compositional characterization (Schwaferts et al., 2019; Laborda et al., 2016). These techniques are particularly useful for well-characterized primary nanoplastics and method-validation or biological studies (Feng et al., 2018; Correia and Loeschner, 2018).

Their application to environmental samples is limited by sampling bias, since only a small fraction of the sample is imaged, as well as extensive sample preparation and potential charging or irradiation artifacts (Schwaferts et al., 2019). Environmental SEM (ESEM) can reduce these limitations by enabling imaging of hydrated or non-conductive samples under low-pressure conditions.

6.3.2.2 Nanoparticle tracking analysis

Nanoparticle tracking analysis (NTA) determines particle number concentration and size distribution by tracking Brownian motion and can detect particles as small as ∼10 nm depending on refractive index (James and Driskell, 2013; Foreman-Ortiz et al., 2022). However, NTA cannot establish whether detected particles are nanoplastics without complementary chemical identification and may have detection limits too high for environmentally relevant concentrations (Zhang et al., 2023; Cai et al., 2021b).

Compared with DLS, NTA is less influenced by large particles in polydisperse samples and has been applied to characterize submicron and nanoplastic particles, including secondary plastics generated by fragmentation (Lambert and Wagner, 2016a; 2016b; Filipe et al., 2010; Gigault et al., 2016). Fluorescence NTA with particle staining can improve selectivity, although interference from sample digestion and staining procedures must be evaluated (Schwaferts et al., 2019).

Dynamic light scattering (DLS) is generally more appropriate for relatively homogeneous fractions obtained after separation techniques such as FFF or CE. Its application directly to complex environmental samples is limited by sensitivity to polydispersity, dust, and aggregates (Schwaferts et al., 2019; Nene et al., 2025).

6.4 Strengths and limits of analytical detection and quantification techniques

Due to the analytical complexity of nanoplastic detection, no single technique suffices for complete characterization across all particle sizes and environments. Thus, major analytical methods vary significantly in the information they provide, detection capability, sensitivity, and limitations. A comparative overview of commonly used techniques is shown in Table 7.

TABLE 7. Comparison of analytical methods used for nanoplastic detection and characterization.

6.5 Analytical detection and quantification challenges: lack of standard methods, reference materials, and harmonized reporting units

Nanoplastics remain among the least standardized contaminants in environmental analysis, with no universally accepted detection and quantification protocol currently available. As a result, exposure estimates and risk assessments often depend on methodological choices rather than true environmental variability (Mitrano et al., 2021). In practical terms, results generated by different laboratories cannot be directly compared due to variation in size thresholds, reporting units, and sample preparation procedures applied across published work (Mitrano et al., 2021).

In addition to instrumental limitations, analytical reliability is strongly influenced by quality assurance and quality control (QA/QC) procedures, including contamination prevention, use of procedural blanks, and validation of recovery rates (Ivleva, 2021; Mitrano et al., 2023). Given the ubiquity of micro- and nanoplastics, airborne contamination during sampling and laboratory processing remains a critical challenge that may significantly affect measured concentrations (Mitrano et al., 2023). Moreover, recovery efficiencies vary considerably depending on extraction and purification protocols, which may introduce systematic bias in reported nanoplastic concentrations (Ivleva, 2021).

Reporting inconsistencies further complicate interpretation. Concentrations are expressed using different metrics, including particle number and mass, which are not directly interchangeable and lack a universally accepted reference unit for comparison (Ivleva, 2021; Mitrano et al., 2023). In practical terms, datasets generated under different reporting schemes cannot be directly integrated without prior methodological alignment, which limits the development of comparable exposure assessments (Mitrano et al., 2023).

The absence of certified reference materials representative of environmentally relevant nanoplastic particles remains a critical limitation. Without such materials, inter-laboratory validation is restricted, and uncertainty associated with measured concentrations cannot be fully constrained (Mitrano et al., 2023). For this reason, progress in nanoplastics analysis depends on the development of standardized methods, representative reference materials, and harmonized reporting practices to enable reliable comparison across studies (Mitrano et al., 2023).

To address these limitations, integrated workflows that combine spectroscopy, thermal analysis, and particle-tracking techniques are essential for detecting various particle sizes. Collaboration among chemists, materials scientists, environmental scientists, and toxicologists is vital for developing standardized protocols to enhance exposure assessment reliability. Current constraints may lead to underestimating nanoplastic levels, creating uncertainty in exposure estimates, so toxicological and risk assessments should be interpreted cautiously. Variability among studies may result from analytical differences rather than true environmental or biological variations. Standardizing protocols, reference materials, and reporting is crucial for reliable data, improved inter-study comparisons, and enhanced risk assessments (Ivleva, 2021; Mitrano et al., 2023).

7 Toxicological evidence

Human evidence on the health risks of micro- and nanoplastics (MNPs) remains limited, with most evidence derived from experimental models (Lamoree et al., 2025; Gouin et al., 2022). The literature is also dominated by polystyrene: 105 of 133 mammalian studies reviewed by J.-L. Xu et al. (Xu et al., 2022) used this polymer. Study quality is another concern; only 10 of 74 oral and two of 74 inhalation studies met all critical criteria in the assessment by Gouin et al. (Gouin et al., 2022). These limitations constrain extrapolation from experimental findings to human health risk (Lamoree et al., 2025; Gouin et al., 2022).

7.1 Particle characteristics, cellular uptake, and barrier interactions

MNP biological behavior depends on polymer type, size, shape, and biocorona (Lamoree et al., 2025). Importantly, cellular uptake does not necessarily result in toxicity. In intestinal models and mice, only a small fraction of 1-, 4-, and 10-µm polystyrene particles were taken up, with no lesions or inflammation observed in vivo (Stock et al., 2019). In Caco-2 cells, exposure to 0.1- and 5-µm particles altered mitochondrial potential and ABC activity, while prolonged exposure caused particle accumulation and ultrastructural changes without substantial genotoxicity or DNA damage (Wu Bing et al., 2019; Domenech et al., 2021). These findings illustrate that particle interaction and uptake should not be equated directly with adverse effects.

7.2 Oxidative stress, inflammation, and genotoxicity

Oxidative stress is among the most consistently reported responses to experimental MNP exposure. Polyethylene and polystyrene particles have been associated with oxidative stress in human cell models (Schirinzi et al., 2017). Polystyrene nanoplastics also induced redox imbalance, inflammation, apoptosis, and reduced transepithelial electrical resistance with depletion of tight-junction proteins in human bronchial and alveolar epithelial models, indicating impaired barrier integrity under the investigated conditions (Yang Ling et al., 2021). In kidney models, polystyrene microplastics were associated with mitochondrial ROS, endoplasmic-reticulum stress, inflammation, autophagy, and histopathological changes (Wang Yung-Li et al., 2021).

Genotoxicity is less consistent. A review of mammalian studies found statistically significant genotoxic effects in 29 of 60 cell-culture results involving polystyrene, without a clear relationship to whether particle size was below or above 100 nm (Møller and Roursgaard, 2023). Conversely, long-term exposure to polystyrene nanoparticles caused cellular accumulation without detectable DNA damage or oxidative stress in Caco-2 cells (Domenech et al., 2021). Thus, oxidative-stress responses are repeatedly observed, whereas genotoxicity appears more dependent on particle characteristics, biological model, and exposure conditions.

7.3 Organ-level and systemic effects in experimental models

Experimental studies indicate that MNPs can interact with intestinal and pulmonary barriers and, under some conditions, reach systemic tissues (Lamoree et al., 2025). For example, orally administered polystyrene microplastics were detected in mouse kidneys and associated with histopathological lesions, cellular stress, inflammation, and altered autophagy-related proteins (Wang Liuwei et al., 2021). However, these effects were demonstrated under specific experimental conditions and cannot yet be quantitatively translated into human health risks because human exposure and dose–response data remain insufficient.

7.4 Reproductive and endocrine effects

Polystyrene microplastics have produced reproductive effects in mouse models. Male mice exposed to 0.5-, 4-, and 10-µm particles for 28 days showed reduced sperm quality and testosterone levels, testicular inflammation, and blood–testis barrier disruption; longer exposure further altered testicular morphology, hormone levels, and sperm parameters (Jin et al., 2021; Jin et al., 2022). Female mice exposed for 35 days showed ovarian inflammation and reduced oocyte quality (Liu et al., 2022). Although these findings indicate potential reproductive toxicity, their relevance to human risk remains uncertain because experimental doses and exposure conditions may not represent typical human exposure (Lamoree et al., 2025).

7.5 Human evidence and translational uncertainty

Direct human evidence currently demonstrates internal exposure more clearly than toxicity. Plastic particles have been detected in blood, placenta, and lung tissue (Ragusa et al., 2021; Amato-Lourenço et al., 2021; Leslie et al., 2022). Emerging clinical studies have reported associations with health outcomes, but these findings remain observational. Patients with inflammatory bowel disease had higher faecal microplastic concentrations than healthy individuals, with concentrations positively correlated with disease severity; however, causality could not be established because disease status may also influence particle retention (Yan et al., 2022). Similarly, MNPs detected in carotid plaques were associated with increased risk of myocardial infarction, stroke, or death during follow-up in a prospective observational study (Marfella et al., 2024). Limited sample sizes and uncertainty in exposure assessment currently restrict causal interpretation and quantitative risk assessment (Lamoree et al., 2025).

Overall, experimental studies provide evidence for oxidative stress, inflammation, barrier disruption, organ effects, and reproductive toxicity under defined laboratory conditions. Figure 8 summarizes the potential toxic effects associated with MP/NP exposure in different biological systems (Malafeev, 2025). Whereas human studies primarily demonstrate internal exposure and emerging associations with clinical outcomes. Differences in particle characteristics, exposure conditions, analytical methods, and study quality remain major barriers to translating experimental findings into human health risk estimates. Representative studies are summarized in Table 8.

TABLE 8. Representative toxicological studies of microplastics and nanoplastics, including major findings and interpretive considerations.

8 Micro/nanoplastics as vectors of co-pollutants

Microplastics and nanoplastics (MNPs) interact with a wide variety of co-occurring environmental contaminants, including heavy metals, polycyclic aromatic hydrocarbons, pharmaceuticals, pesticides, nanoparticles, organohalogens, plastic additives, and organotins (Bhagat et al., 2021). Such interactions can involve aggregation, adsorption, accumulation, transformation, and desorption (Bhagat et al., 2021). If a contaminant is detected on a plastic surface, this does not necessarily imply that plastic is an important biological exposure route. In some exposure situations and for certain contaminants, uptake via plastics may be minor compared with exposure via water and food (Koelmans et al., 2016; Lee et al., 2019).

8.1 Chemical co-pollutants: sorption, transfer, and combined effects

The sorption of antibiotics onto microplastics depends on the physicochemical properties of both the polymer and antibiotic, as well as environmental conditions (Stapleton et al., 2023). Weathering can increase antibiotic sorption capacity by up to 171%, while factors like pH and salinity also influence sorption (Stapleton et al., 2023). Polyamide exhibited the highest adsorption capacity, but in seawater, the capacity was significantly lower compared to freshwater in tests with five antibiotics and microplastic types (Li et al., 2018b).

Particle size can lead to different responses even when the same co-contaminant is studied (Ma et al., 2016). For example, 50-nm nanoplastics significantly enhanced the bioaccumulation of phenanthrene residues in Daphnia magna, while 10-µm microplastics did not affect phenanthrene bioaccumulation, dissipation, or transformation (Ma et al., 2016).

While many studies support the influence of microplastics on bioaccumulation and toxicity, the effects vary by exposure conditions. Sun et al. found that microplastics increased co-contaminant bioaccumulation by 31% and toxicity by 18% across 870 endpoints from 40 studies (Sun et al., 2022). Conversely, a critical review suggested that in ost marine habitats, bioaccumulation from natural prey would likely exceed that from ingested microplastics (Koelmans et al., 2016).

The exposure-context effect has been noted in several experiments (Besseling et al., 2013; Rochman et al., 2013; Lee et al., 2019). Fish exposed to polyethylene with environmental chemicals bioaccumulated these substances, resulting in liver toxicity (Rochman et al., 2013). In contrast, the lugworm Arenicola marina exhibited different results; a polystyrene dose of 0.074% increased PCB bioaccumulation by 1.1–3.6 times, but higher doses showed decreased bioaccumulation, with limited overall effects (Besseling et al., 2013). A fish bioaccumulation model estimated that uptake of hexachlorocyclohexanes via ingested polyethylene was negligible compared to uptake via water and food (Lee et al., 2019). Sorption alone is not enough to establish an environmentally important vector effect. Plastic-mediated exposure depends on contaminant properties, particle characteristics, and alternative exposure pathways under studied conditions (Koelmans et al., 2016; Lee et al., 2019; Sun et al., 2022).

8.2 Pathogens and antibiotic resistance genes

Plastic surfaces can host microorganisms, including pathogens like Vibrio parahaemolyticus, found on microplastics in the North and Baltic Seas (Kirstein et al., 2016a).

Antibiotic resistance genes (ARGs) have also been found to be enriched in the microplastisphere, but the degree of enrichment is highly dependent on the material used for comparison (Yu et al., 2022). A meta-analysis of 477 paired ARG-abundance measurements representing 26 ARG types found that ARG abundance was greater on microplastics than on inorganic substrates and in surrounding natural water. ARG abundance on microplastics was nevertheless lower than that measured on natural organic substrates (Yu et al., 2022). Experimental conjugative-transfer studies provide further evidence: under the investigated conditions, microplastic exposure increased bacterial conjugative transfer frequency by approximately 1.4–1.7-fold (Yu et al., 2023).

The presence of potentially pathogenic microorganisms on plastic surfaces does not confirm transmission or infection. Few studies show direct evidence of microplastic-mediated infection, and key factors like viability, detachment, and dose delivered to a host are often unmeasured. The relevance of pathogen transport depends on the viability of microorganisms on plastic (Ahmad et al., 2026).

Current evidence suggests that, under specific conditions, chemical sorption occurs, leading to increased contaminant bioaccumulation and toxicity in some systems, enrichment of ARGs, and enhanced horizontal gene transfer. The magnitude and relevance of these effects vary by contaminants, particles, biological systems, and exposure scenarios (Koelmans et al., 2016; Sun et al., 2022; Yu et al., 2022; Ahmad et al., 2026). Table 9 summarizes key studies on microplastic- and nanoplastic-mediated interactions with co-pollutants.

TABLE 9. Representative studies of microplastic- and nanoplastic-mediated interactions with co-pollutants.

9 Micro- and nanoplastics risk assessment

Risk assessment for micro- and nanoplastics (MNPs) involves integrating exposure, hazard information, and particle characteristics while considering assessment uncertainties (Noventa et al., 2021; Koelmans et al., 2022). This is complicated by the variability in particle size, shape, density, and chemical properties, which affects the correlation between exposure and effect data (Koelmans et al., 2022). Proposed frameworks for environmental and human-health risk assessment aim to tackle these challenges (Gouin et al., 2019; Noventa et al., 2021; Koelmans et al., 2026). Recent comparisons of human-health frameworks show variations in detail, complexity, application, and implementation stages (Koelmans et al., 2026).

9.1 Risk-assessment frameworks and particle-specific considerations

The Human Health Risk Assessment Framework for Nano- and Microplastics (HRA-NMP) maintains traditional elements like hazard identification, characterisation, exposure assessment, and risk characterisation (Noventa et al., 2021). It distinguishes physical, chemical, and microbiological hazards and incorporates particle-specific factors into the process. The framework combines analytical techniques, empirical data, theoretical and modelling approaches, and stakeholder engagement to overcome certain limitations of MNP assessment (Noventa et al., 2021). Other methods suggest representing particle diversity with probability density functions (PDFs), which help align exposure and effect thresholds, allowing continuous particle properties to be part of the assessment instead of treating microplastics as homogeneous or disconnected categories (Koelmans et al., 2022). Not all human-health frameworks are equally developed; recent assessments show substantial variation in tier level, detail, complexity, target users, endpoints, and implementation, with many components still under development and not yet fully regulatory (Koelmans et al., 2026).

9.2 Environmental risk assessment

Exposure and effects data are limited, constraining environmental risk assessment of microplastics (Gouin et al., 2019). A review highlighted the need for better data and standardised reference materials to improve assessment. Adam et al. conducted a preliminary freshwater risk assessment using exposure data from Asia, Europe, and North America, with effect data in a probabilistic species sensitivity distribution. Most risk ratios were below one, except in Asia where 0.4% exceeded one (Adam et al., 2018). The assessment was deemed preliminary due to dependence on data quality and homogeneity.

A risk-based management approach utilized species sensitivity distributions and data from 21 peer-reviewed microplastic toxicity studies (Mehinto et al., 2022). Expert evaluation showed high confidence in the overall framework, but individual threshold confidence varied. In 2025, two mathematical errors were found, and one study was excluded for not meeting quality criteria, leading to revised threshold values. The corrected values remained within previous confidence intervals, and the overall conclusions remained unchanged (Mehinto et al., 2025).

Uncertainty in exposure information can impact risk characterization, as seen in the San Francisco Bay study. The highest relative uncertainty came from correcting fibre data in manta-trawl monitoring. Thus, the quality, completeness, and alignment of monitoring data are crucial for deriving threshold-based environmental risk estimates (Coffin et al., 2022).

9.3 Human-health risk assessment

Selecting an assessment framework is not the only challenge for human health; reliable exposure and hazard methods are limited, and datasets are heterogeneous (Vogel et al., 2024; Lane et al., 2025). The POLYRISK framework, designed as a modular system focusing on inhalation exposure, is structured around physicochemical characterization, inhalation potential evaluation, and toxicological assessment (Vogel et al., 2024). It includes integrated approaches like IATAs, NAMs, and grouping approaches where relevant.

Exposure estimates involve variables like population traits, exposure pathways, activity factors, particle features, and uncertainty. Lane et al. developed demographic-specific scenarios for nine groups—adults, women, seniors, diseased individuals, high-hazard workers, and children—and recommend probabilistic methods to address variability and uncertainty in exposure data (Lane et al., 2025).

An alternative modular approach using IATAs, adverse outcome pathways (AOPs), and NAMs has been developed for oral exposure. It is composed of three phases: physicochemical characterisation of the test material, assessment of intestinal crossing, and hazard-assessment testing (Vogel et al., 2026).

The availability of multiple frameworks does not indicate full maturity in human-health risk assessment. Koelmans et al. compared six frameworks, noting differences in detail, complexity, purpose, and readiness. Some components are still under development or need validation before regulatory use (Koelmans et al., 2026).

Table 10 outlines key frameworks and applications for microplastic and nanoplastic risk assessment, highlighting advancements in integrating particle-specific features and exposure variability. However, differences in scope, data requirements, and implementation reveal ongoing uncertainties in foundational evidence. These uncertainties range from environmental fate and analytical detection to hazard evaluation, impacting the accuracy and comparability of risk estimates.

TABLE 10. Representative frameworks and applications for microplastic and nanoplastic risk assessment.

9.4 Critical synthesis, key evidence gaps, and uncertainties

Environmental transformation involves aging that changes particle size and surface properties (Lu et al., 2023), while particle fate is influenced by physicochemical and biological processes (Jacquin et al., 2019; Richard et al., 2019; Gong et al., 2019). These changes impact the transport and biological interactions of microplastics (MPs) and nanoplastics (NPs). While the small size of NPs enhances their mobility, environmental conditions play a significant role in this behavior (Alimi et al., 2018; Chen et al., 2023). Increased ionic strength, organic matter, or biofilm formation can lead to aggregation, increasing particle size and density, which promotes sedimentation (Shiu et al., 2020). Therefore, differences in NP mobility and sediment accumulation should be viewed in the context of environmental chemistry and aggregation processes rather than suggesting different transport mechanisms.

The formation and persistence of plastic particles are significantly influenced by environmental conditions. While weathering contributes to fragmentation, its quantitative impact on nanoplastic (NP) production is challenging to measure due to variations in polymer composition, crystallinity, additives, and exposure conditions (Burrows et al., 2024; Rostampour et al., 2025). Degradation rates also differ widely based on polymer type and setting, with biodegradable polymers degrading rapidly in enriched or thermophilic environments, unlike chemically resistant ones (Pires et al., 2024). Definitions of degradation vary, complicating study comparisons. Although the sorption of organic contaminants and metals onto plastics is well established, this does not necessarily enhance their bioavailability or toxicity, as plastics also interact with sediments and organic matter that can act as contaminant reservoirs (Martín et al., 2022; Cimpean et al., 2026). Thus, environmental transformations and contaminant interactions must be understood in the context of their surrounding matrix.

Toxicological evidence indicates reliance on particle characteristics and experimental conditions, with oxidative stress, inflammation, membrane interactions, and cellular uptake showing variable magnitude across systems (Thornton et al., 2022; Hossain and Sahajwalla, 2026). Variability arises partly from using pristine, commercial spherical particles at often high, environmentally irrelevant concentrations (Elert et al., 2017; Lenz et al., 2016). These particles are valuable for isolating variables like size, shape, surface functionality, and establishing mechanistic baselines (Gouda et al., 2025). However, they do not fully reproduce the irregular morphology, surface modification, biofouling, aggregation, and chemical complexity of aged particles. Therefore, differences in toxicological effects may stem from biological responses, particle properties, exposure conditions, or experimental design.

Methodological limitations constrain interpreting evidence. Standardized protocols for assessing MNP hazards and ecotoxicity are limited (Abdolahpur Monikh et al., 2023; Boccia et al., 2024). Exposure methods for soluble chemicals may not suit insoluble particles, which aggregate and settle based on pH and ionic strength (Abdolahpur Monikh et al., 2023). Differentiating plastic effects from background particulates is difficult because natural waters contain many non-plastic suspended solids (Gerdes et al., 2019). Variability in sample prep, particle recovery, detection, characterization, and the absence of standardized methods, reference materials, QA/QC, and reporting units hinder result comparability (Gouin et al., 2022).

These uncertainties limit environmental risk characterization. While research on MPs has progressed, formal risk assessments face major gaps in occurrence, exposure, and hazard data (Vighi et al., 2021). For NPs, these gaps are more significant due to the challenges in detecting, isolating, and quantifying them in complex environmental matrices (Asselman, 2021; Cunningham et al., 2023). The main evidence gap exists in connecting environmental transformation and fate with reliable analytical detection, accurate exposure characterization, and ecologically relevant hazard assessment. Addressing these uncertainties is crucial for developing scientifically robust and realistic risk estimates.

10 Future perspectives and research needs

Despite rapid advances in microplastic and nanoplastic (MNP) research, important methodological and knowledge gaps continue to limit reliable environmental and human-health risk assessment. Four priorities should guide future research.

First, prioritize analytical standardization by developing and validating harmonized methods for sampling, detection, and quantification of nanoplastics. This effort should include the use of certified reference materials and consistent reporting metrics to ensure that reported concentrations reflect actual contamination levels rather than methodological variability.

Second, there is an urgent need for relevant exposure data. Current estimates are constrained by limited monitoring and varying analytical methods, particularly for smaller particles. Coordinated monitoring across environmental compartments, along with consistent metrics, will help clarify exposure ranges and reduce uncertainties in health assessments.

Third, toxicological research should align more closely with realistic environmental conditions. Future studies must consider aged particles, chronic low-dose exposure, bioavailability, and interactions with other contaminants, rather than focusing solely on pristine samples and short-term exposures.

Finally, these improvements should be integrated into harmonized risk-assessment frameworks. Future assessments must link particle transformations and environmental fate to detection, exposure, biological effects, and uncertainties. This integration will help distinguish certain hazards from uncertain effects, ultimately strengthening the scientific foundation for regulatory decision-making.

11 Conclusion

Microplastics and nanoplastics (MNPs) are dynamic environmental contaminants whose distribution, persistence, and biological interactions are governed by interconnected processes of transport, transformation, aggregation, and biological modification. Environmental aging, fragmentation, biofilm formation, particle size, surface chemistry, and surrounding environmental conditions can substantially alter their mobility, environmental fate, and potential interactions with organisms and co-occurring contaminants.

A major finding of this review is that uncertainty in MNP risk assessment arises not only from incomplete knowledge of toxicity but also from limitations in detecting, characterizing, and quantifying environmentally relevant particles. This is particularly important for nanoplastics, for which differences in operational size definitions, analytical detection capabilities, and reporting approaches can affect measured concentrations and complicate comparisons among studies.

The toxicological literature provides evidence for biological responses including oxidative stress, inflammation, and cellular effects; however, the consistency and environmental significance of these responses remain uncertain. Differences in particle characteristics, aging state, exposure concentrations, experimental conditions, and biological models contribute to variability among studies. In particular, the frequent use of pristine particles and exposure concentrations or scenarios that may not represent environmental conditions limits direct extrapolation of laboratory findings to real-world risk.

Overall, the evidence reviewed indicates that MNP risk cannot be reliably evaluated by considering environmental occurrence, particle transformation, analytical detection, exposure, or toxicity in isolation. These components form a connected continuum in which environmental transformation can alter particle characteristics and detection, while analytical limitations influence exposure estimates and toxicological interpretation. Addressing these interdependencies through standardized analytical approaches, environmentally relevant exposure assessment, realistic toxicological models, and explicit consideration of uncertainty will be essential for determining the actual significance of MNP contamination and establishing scientifically robust risk assessments.

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Footnotes

Edited by: Mazeyar Parvinzadeh Gashti, Pittsburg State University, United States

Reviewed by: Bulgariu Laura, Gheorghe Asachi Technical University of Iaşi, Romania

Eleonora Nicolai, Saint Camillus International University of Health and Medical Sciences, Italy