Chemical reactivity of weathered nanoplastics and their interactions with heavy metals

Abstract

There is growing concern about the threat that nanoplastics (NPs) pose to ecosystems. However, a comprehensive risk assessment of NPs is currently constrained by the paucity of knowledge on the chemical reactivity of NPs, which were previously thought to be chemically inert. This review identifies the chemical reactivity of NPs that have undergone abiotic and biotic weathering, including the formation of free radicals, the increase in the content of oxygen-containing functional groups, and the release of plastic leachates. Their interaction with legacy contaminants, such as heavy metals (HMs), is then examined, highlighting their critical role in the oxidation and reduction of HMs, through free radical-mediated redox processes and electron shuttling by carbonyl groups. This review offers new insights into the risk of NPs, where their interaction with legacy contaminants determines the long-term exposure scenario for ecosystems. The unexpectedly large pool of reactive NPs in nature will not only affect their risks but also impact the biogeochemistry of HMs and other contaminants that could react with free radicals and carbonyl groups.

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Environmental significance

Nanoplastics (NPs) are ubiquitous and an emerging concern in the natural environment. The chemical reactivity of NPs has significant implications for their risks to human and ecosystem health. This review provides a comprehensive understanding of the chemical reactivity of NPs that have undergone abiotic and biotic weathering. We suggest that these reactive NPs are capable of triggering the transformation of legacy contaminants, such as heavy metals. The identification of the chemical reactivity of NPs offers new insights into their role in the geochemical cycling of heavy metals and the associated risks.

1. Introduction

The rapid growth of plastic production, coupled with inefficient disposal and recycling systems, has resulted in the release of plastic waste into the environment.^1–3 The accumulated plastics in the environment are subject to dynamic weathering via biotic (e.g., mediated by microorganisms) and/or abiotic processes (e.g., induced by ultraviolet light), resulting in the fragmentation of plastics into nanoplastics (NPs, <1000 nm).^4–8 The presence of these NPs has been confirmed in a range of environmental samples. For instance, using pyrolysis gas chromatography-mass spectrometry (Py-GC-MS), Ter Halle et al. estimated that the spectrometric signal of the colloidal aromatic fraction (<1000 nm) of seawater in the North Atlantic subtropical gyre was attributable to a mixture comprising 73% polyvinyl chloride (PVC), 18% polyethylene glycol terephthalate (PET), and 9% polystyrene (PS).⁹ Wahl et al. observed polyethylene (PE), PS, and PVC within a 20–150 nm size range in water extracts from soil.¹⁰ Additionally, Materić et al. identified PET NPs at 5.4–27.4 ng mL⁻¹ in surface snow samples from the Australian Alps.¹¹ Despite these findings, the analysis of NPs in environmental samples is still emerging, with current protocols immature and needing further development. It is estimated that NPs are generally not intentionally designed and vary widely in shape, size, polydispersity, additives, adsorbed contaminants, surface properties, and composition due to different source materials, fragmentation pathways, and environmental exposure.⁶ The resulting physical and chemical heterogeneity of NPs may influence their reactivity and will certainly affect their interactions with natural components and organisms.⁶ Pieces of evidence show that NPs can interact with and cross biological barriers, including epithelial tissues and cell membranes,^12,13 posing toxicity risks to organisms. Reported impacts include oxidative stresses, immobility, disrupted organ function, reduced reproduction, and potential bioaccumulation throughout food chains.^14–17 However, biased risk assessments of NPs could be obtained by ignoring the coexistence of NPs and legacy contaminants, such as heavy metals (HMs), in nature.

NPs and HMs co-occur either through the incorporation of HMs into NP manufacturing or through their common pathways to the environment. In particular, the improper handling of plastic products and the consequent escalating utilization of these materials has resulted in a substantial influx of NPs into the environment.¹⁸ The majority of HMs (e.g., Cd, Pb, Zn, Hg) serves a variety of functions in the production of plastic, acting as pigments, stabilizers, and fillers.^19–21 These HMs are readily released following weathering processes. For example, 26.5% of Pb in PVC was released upon UV irradiation for 40 h,²² while 62.2% of Ni, 6.5% of Cd, and a range of 5.4–40.3% of Zn present in electronic plastics were released upon natural light irradiation for 112 days.²³ In parallel, NPs inevitably interact with HMs in municipal wastewater plants, surface waters, sediments, and soils.^24,25 The highly likely interaction between NPs and HMs may introduce an uncharacterized variable in the exposure of NPs, which results in toxicological effects that differ from those observed in NP exposure alone.

There is no consensus on the toxicological effects of NPs in the presence of HMs. Nanoplastics (e.g., PS at 300–9000 nm, PE at 50–533 nm) and HMs (e.g., Ag, Au, Cu, Zn, Pb, Cd) have been reported to exert synergistic toxic effects on a variety of organisms, including the lettuce Lactuca sativa L., the alga Chlorella vulgaris, Daphnia magna, and zebrafish embryos.^26–29 Elsewhere, antagonistic interaction has been reported between NPs and HMs. For example, Alaraby et al. demonstrated that PS NPs (80 nm) reduced oxidative stress and genotoxicity in Drosophila in the presence of Ag(I), compared to the Ag(I) group alone.³⁰ The discrepancy in the literature can be attributed to the differences in the physicochemical properties of NPs (e.g., size and type) and biological traits.^31–33 Moreover, it is most likely that the interaction between NPs and HMs is not considered, where PS could induce the oxidation of Ag(0) to Ag(I).³⁴ It is therefore crucial to examine the chemical reactivity of NPs and their interaction with HMs.

To gain insight into the interaction between NPs and HMs, we analyzed the literature of the past 34 years from the Web of Science database using Citespace (Fig. 1). A total of 263 publications were identified using the keywords “Nanoplastic* OR Nano plastic*” and “Heavy metals*”. The first publication on NPs and HMs appeared in the 1990s, focusing on the application of HMs in plastic production. From 2000 to 2010, most studies focused on the properties of plastics. In the past 10 years, there have been efforts to examine the absorption of HMs on NPs, and their combined effects. However, the chemical reactivity of NPs remains less understood.

Fig. 1 Evolution of the field of the interaction between NPs and HMs and co-occurrence of keywords. All data used were retrieved from Web of Science (WOS) from 1990 to 2024. The font size of topic description text is proportional to the number of publications.

This review begins by discussing the weathering and the resulting chemical reactivity of NPs. It then examines the role of weathered NPs in the adsorption and transformation of HMs. Finally, we identify areas requiring further research and suggest potential avenues for future studies. Understanding the chemical reactivity of weathered NPs, and thus their interaction with HMs, is crucial for determining the long-term exposure scenario for ecosystems as well as for developing a comprehensive risk assessment of NPs.

2. Chemical reactivity of weathered NPs

Weathered NPs undergo physicochemical modification, leading to the chemical reactivity, including the production of free radicals, an increase in the content of oxygen-containing functional groups, and the release of plastic leachates (Fig. 2).

Fig. 2 The chemical reactivity of NPs, including the formation of free radicals, the increase in the content of oxygen-containing functional groups, and the release of plastic leachates. (R·, tertiary alkyl radical; ROO·, tertiary peroxy radical; ROOH, hydroperoxide group; RO·, tertiary alkoxy radical; PAOR, persistent aminoxyl radical; SR·, sulfonyl radical).

2.1 The generation of free radicals from weathered NPs

2.1.1 Weathered NPs upon light irradiation generates free radicals. Free radicals are produced from weathered NPs when exposed to light.^35–38 Sunlight is mainly composed of infrared (wavelength λ = 700 nm to 1 mm), visible (λ = 400–700 nm), and UV (λ = 100–400 nm) light.³⁹ The UV fraction of sunlight plays a critical role in the weathering of NPs⁴⁰ because it has sufficient energy (299–412 kJ mol⁻¹ for 290–400 nm) to cleave the C–C bond (284–368 kJ mol⁻¹) and the C–H bond (381–410 kJ mol⁻¹) in aromatic plastic polymers such as PS.⁴¹ The cleavage of the C–H bond could induce the formation of tertiary alkyl radicals (R·), which can subsequently react with atmospheric oxygen to form tertiary peroxy radicals (ROO·).⁴² These tertiary peroxy radicals undergo reactions involving the abstraction of hydrogen to form hydroperoxide groups (ROOH), which are inherently unstable and photolabile.⁴³ As shown by Ranby et al., the dissociation energies of CO–OH, C–OOH, and COO–H are 175.8, 293.0, and 376.7 kJ mol⁻¹, respectively, which fall within the energy range of UV radiation (299–412 kJ mol⁻¹).⁴⁴ This suggests that hydroperoxide groups would be cleaved to form tertiary alkoxy radicals (RO·) and reactive oxygen species (ROS), including hydroxyl radicals (·OH), superoxide (O₂˙⁻) and H₂O₂.⁴⁴ These ROS, which are highly reactive, can attack plastic polymers and further induce cleavage of C–C and C–H bonds, thereby accelerating the weathering of plastic polymers.^42,45 As for aromatic plastic polymers containing the O–H bond, such as phenol-formaldehyde resin, cleavage can occur at the O–H bond attached to the benzene ring, resulting in the formation of phenoxyl free radicals. These then react with atmospheric oxygen to form ROS.⁴² In addition, polymers containing nitrogen or sulfur can produce other free radicals, including those centered on atoms other than carbon or oxygen. For example, light-weathered non-aromatic polyamides (PAs) generated reactive nitrogen species (RNSs), such as ·NO₂.⁴⁶ This was attributed to the generation of amine oxide moieties in light-weathered PAs, which served as precursors for the formation of persistent aminoxyl radicals (PAORs). These radicals subsequently led to the formation of reactive nitrogen species (RNSs) through a reaction with peroxide radicals and O₂˙⁻.⁴⁶ In addition, light-weathered poly(phenylene sulfide) (PPS) produced reactive sulfur species (SO₃˙⁻),⁴⁷ because the cleavage of the C–S bond in PPS during light weathering led to the formation of sulfonyl radicals (SR·), which reacted with ·OH to form SO₃˙⁻.⁴⁷

Most studies focus on sunlit surfaces, with little attention to non-sunlit environments. In fact, the majority of NPs persist in the dark environments, including deeper water or landfills.^48,49 Whether and how free radicals would be produced during light-independent weathering remains largely unknown. Chen et al. have demonstrated that semiquinone radicals and H₂O₂ were produced from phenol-formaldehyde resins after hydrogen peroxide treatment in the dark.⁵⁰ Future research on light-independent weathering of NPs is urgently needed.

2.1.2 Chemical reactivity of light-weathered NPs. Light-weathered NPs are redox reactive, which is associated with the formation of free radicals.⁵¹ For example, ·OH is highly reactive, with a higher oxidation potential of 2.8 V, which can attack the C–H bond of a wide range of molecules.⁵² Light-weathered PS NPs (10–1000 nm) with these ·OH can trigger abiotic transformation of HMs through oxidation processes, such as the oxidation of Ag(0) by the production of ·OH.³⁴

The generation of free radicals from light-weathered NPs can be quantified using electron paramagnetic resonance (EPR) or chemical probe methods. After 60 days of light weathering, Li et al. used EPR to demonstrate that the levels of O₂˙⁻, ·OH, and ¹O₂ produced from PS NPs (100 nm) increased by 8.6-, 15.3-, and 4.3-fold, respectively.⁵³ Using chemical probe methods, Zhu et al. demonstrated that light-weathered PS (0.05 g) can generate O₂˙⁻ and ·OH at concentrations of 0.003 and 0.0003 μmol g⁻¹ d⁻¹, respectively.⁵⁴ It is thus estimated that 500 g of light-weathered PS could induce O₂˙⁻ formation at a level of 1.5 μmol d⁻¹. These concentrations of O₂˙⁻ are comparable to those detected in floodplain, riverine wetland, and river band soils (0.3–4 μmol d⁻¹).⁵⁵ Collectively, weathered NPs play a previously unrecognized role in contributing to total ROS production in nature, with significant implications for biogeochemical processes of elements.

2.2 The increase in oxygen-containing functional groups on weathered NPs

2.2.1 The oxidation of NPs through abiotic and biotic weathering. Abiotic pathways, including light weathering which induces oxidation through free radical reactions,³⁶ mechanical strength, alkaline conditions, heat, and chemical reagents, contribute to the oxidation of NPs.^56–58 Abiotic weathering initially occurs on the surface of NPs, resulting in the formation of cracks and pits.^56,57 This process allows oxygen access to the inner layer, leading to the oxidation of NPs.⁵⁶ The carbonyl index (CI), which normalizes the height or area of the carbonyl (C=O) peak to a stable reference peak, has been utilized to measure the extent of polymer oxidation.⁴⁶ For example, using the CI values as a measure, Luo et al. demonstrated that the CI values of PE increased rapidly in the initial stage of 2 weeks during thermal weathering but remained stable after 5 weeks.⁵⁸ Additionally, Yu et al. observed a significant positive correlation between the CI values of PE and UV irradiation time, indicating that oxidation could be effectively predicted.⁵⁹

Abiotic oxidation of NPs is influenced by natural organic matter (NOM) and clay minerals. There is no consensus on the role of NOM in the oxidation of NPs. Suwannee River fulvic acid (SRFA) has been reported to facilitate the oxidation PS NPs (50–1000 nm) by generation of ·OH and O₂˙⁻ under light irradiation, resulting in an 8–11-fold increase in O/C ratios.⁶⁰ In contrast, Cao et al. demonstrated that SRFA quenched ·OH and weakened the oxidation process, and the oxygen-containing functional groups did not vary significantly on PS NPs (100 nm).⁶¹ These disparate results could be attributed to the heterogeneity of PS NPs, which can be oxidized rather through laboratory weathering or during the polymer processing and/or storage. The latter was demonstrated by a small amount of oxygen atoms (2.21%) and oxygen-containing functional groups observed on unweathered PS.⁴² We have observed oxygen-containing functional groups in PS NPs (1000 nm) either before or after laboratory weathering under stimulated sunlight (Fig. 3a, unpublished data). Minerals are another important factor. Kaolinite and montmorillonite have been demonstrated to accelerate the oxidation of PVC and PET via the formation of ·OH, resulting in a 1.5–1.8-fold increase in the O/C ratio under light irradiation.⁶² Similar phenomena have also been observed with iron (hydr)oxides, whose O/C ratio increased by 1.6–2.2-fold under irradiation.⁶³ Collectively, these studies demonstrate the impacts of NOM and clays on the oxidation of NPs under irradiation, with less consideration of other weathering processes.

Fig. 3 Oxygen-containing functional groups (a) and free radicals (b) can be formed on PS NPs (1000 nm) either through laboratory weathering or during the polymer processing and/or storage. The PS NPs in (a) and (b) were sourced from different manufacturers.

In addition to abiotic pathways, biotic weathering can involve the oxidation of NPs by the formation of eco-corona, microbial degradation of NPs, and grinding NPs by animals. The biomolecular coating that often forms on NPs is often referred to as the eco-corona.⁶⁴ This coating includes proteins as well as a diverse array of other biomolecules such as metabolites from cellular activity and/or NOM.⁶⁴ The biomolecular coating is rich in functional groups, which provide oxygen-containing functional groups on NPs. For example, Zhu et al. demonstrated that the O/C ratio increased by 0.7–1.4 times and 0.3–0.8 times, respectively, with the formation of the eco-corona on weathered PS NPs (448 nm) and weathered HOOC-PS (476 nm).⁶⁵

These biomolecules can trigger the degradation of NPs, which involves the oxidation of NPs. Enzymes derived from multiple microorganisms such as Bacillus subtilis, Thermobifida fusca, and Pseudomonas sp. have been demonstrated to degrade plastics, such as PET and polyurethane (PUR), primarily through hydrolysis of ester bonds.^5,66–68 In the case of non-hydrolyzable polymers, such as PE and PS, enzymes degrade them through the oxidation occurring in-chain and at the termini, resulting in the formation of dicarboxylic acids, alcohols, fatty acids, ketones, aldehydes, and esters.⁶⁹ Using Fourier-transform infrared (FTIR) spectroscopy, Gao et al. detected the formation of carbonyl groups on PE following the incubation with A. alternata FB1.⁷⁰ Furthermore, Sathiyabama et al. demonstrated that lignolytic enzymes derived from Cladosporium sphaerospermum increased the content of carbonyl groups in weathered PE by 1.0–1.3 times compared to the unweathered counterparts.⁷¹ In addition to enzymes, biotic secretion may contribute to the oxidation of NPs. Using in situ FTIR, Zhang et al. demonstrated that PS NPs (20–1000 nm) incubated with secretions derived from amoeba D. discoideum exhibited a unique carbonyl group stretching vibration peak at 1730–1750 cm⁻¹.⁷² In terms of NOM, they are capable of participating in the oxidation of PS NPs (1000 nm) on both sunlit surfaces and in the absence of light.⁷³ This can be achieved through the formation of ROS, including O₂˙⁻, ·OH, and ¹O₂, resulting in an increase of up to 10.7-fold in the O/C ratio in PS.⁷³ Given that NPs are smaller than microorganisms, oxidation through colonization of microorganisms is unlikely to occur in NPs.⁷⁴ Our current knowledge about microbial interactions with NPs is limited by the challenge of efficient extraction and analysis of NPs in environmental and biological matrices.⁷⁵

Another consideration is grinding NPs by animals, rather than through the well-documented microbial degradation that occurs in their guts.⁷⁶ Using confocal Raman microscopy (LCM-Raman) and scanning electron microscopy (SEM), Zhao et al. demonstrated that rotifers can grind the edges of PS, with irregular tear marks on weathered PS.⁷⁷ These tear marks enable accessibility of oxygen to the internal layer of plastics, promoting the oxidation.⁵⁶ Other animals species have been reported to grind plastics, such as Antarctic krill (Euphausia superba),⁷⁸ yet there is a paucity of knowledge regarding the impact of such action on the degree of oxidation of NPs.

2.2.2 The chemical reactivity of oxidized NPs. The oxidized NPs are redox reactive due to their oxygen-containing functional groups, including carbonyl (C=O), carboxyl (–COOH), hydroxyl groups (O–H), and quinones.^45,79,80 For example, oxidized phenol-formaldehyde resins had electron-donating capacity (0.264–1.15 mmol e⁻ g⁻¹) and electron-accepting capacity (0.120–0.300 mmol e⁻ g⁻¹) due to the quinone and carboxyl groups.⁵⁰ These functional groups have been documented to facilitate the transformation of HMs, such as the reduction of Ag(I) and Cr(VI).^81,82 Additionally, oxidized PS had electron-rich carbonyl groups, which were responsible for the transformation of HMs via complexing with charged HMs followed by electron shutting.⁸³

The quantification of oxygen-containing functional groups on oxidized NPs can be achieved through the application of Boehm titration. The contents of carboxyl groups and phenolic hydroxyl on oxidized PVC were 0.15 mmol g⁻¹ and 0.18 mmol g⁻¹, respectively.⁸⁴ These values can be approximately 1.9–7.5 times and 3.6–18 times greater than those observed in other carbon-containing materials, such as biochars, respectively.⁸⁵ The functional groups on biochars play a role in contaminant transformation, either through direct electron transfer or as an electron shuttle between bacteria and elements.^86,87 It is thus suggested that oxidized NPs can contribute significantly to the contaminant transformation, particularly in plastic-rich systems. Moreover, NPs are composed of carbon but also include other elements such as nitrogen in PA, chlorine in PVC, and sulfur in PPS. The capability of these NPs to produce redox-reactive functional groups remains less understood.

2.3 The release of plastic leachates

Plastic weathering triggers the release of leachates containing additives or organic substances derived from polymer degradation.^88–90 Sheridan et al. showed that PE leachates were released at a level of 0.1 mg C per L into the freshwater in the field.⁹¹ Furthermore, accelerated weathering of PE in the laboratory resulted in the release of leachates at concentrations of 2.4–5.6 mg C per L.^92,93 In this study, the leachates from weathered NPs are categorized into plastic additives and plastic oligomers.

2.3.1 Plastic additives in leachates. More than 10 000 additives, such as phthalic esters and flame retardants, have been incorporated into plastics.⁹⁴ In some cases, additives make up as much as 70% of the total weight of the plastic.⁹⁵ These substances are physically bound to polymers and feasibly released into the environment during the plastic life cycle. In the field, plastic film released 11.3 tonnes of di-n-butyl phthalate and 25.3 tonnes of ethylhexyl phthalate into the soil, respectively.⁹⁶ Additionally, PE can release 94.2% of tris(2,4-di-tert-butylphenyl) phosphate in the soil slurry within 2 days.⁹⁷

These additives are chemically reactive. For example, bisphenol A, a synthetic compound used in plastics and resins, can complex with Cr(VI) via its electron-rich groups.⁹⁸ Additionally, our preliminary experiments have demonstrated that 2,6-di-tert-butylphenol, a common antioxidant used in plastics, facilitated the reduction of Ag(I) to Ag(0) (unpublished data). However, the chemical reactivity of additives that are released from weathered NPs remains largely unknown. A major challenge is the identification of these large amounts of additives in environmental and biological matrices.⁹⁹ To address this, non-targeted analysis based on high-resolution mass spectrometry and computational tools will be an ideal approach to identify the chemicals in additives released from NPs.¹⁰⁰ For example, using non-targeted analysis, Tang et al. identified 674 formulas of halogenated organic additives in the plastics of three electrical products. Of these, 166, 362, and 146 were identified as organochlorine, organobromine, and mixed chlorinated/brominated organic compounds, respectively.¹⁰¹ It is therefore necessary to differentiate the chemical reactivity of numerous additives present in the environments.

2.3.2 Plastic oligomers in leachates. Plastic oligomers are by-products of polymer synthesis and degradation intermediates, consisting of chains of 2 to 40 repeating monomer units.¹⁰² They have molecular weights ranging between 10⁴ and 10⁶ Da.¹⁰² Plastic oligomers can result from the incomplete polymerization of monomers and additives, known as “primary oligomers”.^103,104 Approximately 1% of the total weight of plastic polymers can be attributed to these primary oligomers.^105,106 In addition, plastic fragmentation and weathering can result in the formation of oligomers, where high-degree oligomers undergo a transformation into lower-degree derivatives known as “secondary oligomers”.¹⁰² For example, Wang et al. demonstrated that enzymatic hydrolysis of polylactic acid (PLA) generated oligomers during gastrointestinal processes.¹⁰⁷

While the bioavailability and toxicity of plastic oligomers have been documented,^107,108 their chemical reactivity remains unclear due to the challenges of characterization methodologies. Kwon et al. used benzene and dichloromethane to extract styrene oligomers from sand and seawater, respectively,¹⁰⁹ while Yang et al. utilized ethanol to extract water-insoluble PET oligomer agglomerates.¹¹⁰ These analytical methods, however, cannot target all the oligomers released from plastic polymers. The structural diversity of oligomers results in a wide range of physical and chemical properties (e.g., solubility), even among oligomers derived from the same polymer. For example, 12-mer derived from PLA is hydrophobic, whereas the dimer derived from PLA is hydrophilic.¹⁰² Thus, ad hoc synthesis of oligomer standards with different chain lengths and stereochemistry is essential to analyze the oligomers in environmental and biological samples.

3. Interaction between weathered NPs and HMs

3.1 Adsorption of HMs on weathered NPs

The aforementioned evidence demonstrates that weathered NPs are chemically reactive, which can facilitate the adsorption of HMs. The adsorptive interaction between weathered NPs and HMs has been extensively studied and reviewed.^{36,40,111,112} Generally, HMs are adsorbed onto weathered NPs via physical (e.g., pore filling and electrostatic interaction) and chemical mechanisms (e.g., surface complexation, hydrogen bonding, and cation–π interaction) (Fig. 4a).

Fig. 4 Interaction between weathered NPs and HMs. (a) Adsorption of HMs on weathered NPs. (b) Transformation of HMs by weathered NPs.

Generated pores on weathered NPs can increase the adsorption of HMs compared to their unweathered counterparts. This can be attributed to the increase in the specific surface area and pore volume of weathered NPs. For example, SEM analysis demonstrated that light-weathered PS (100 nm) created surface pores and thus enhanced the adsorption of Pb(II), Cu(II), Cd(II), Ni(II), and Zn(II).¹¹³ In addition, weathering can increase the surface charges of NPs through the formation of functional groups such as carboxyl acids, thereby enhancing their electrostatic interaction with HMs.¹¹⁴ For example, light-weathered PP showed a 76% reduction in zeta potential compared to its unweathered counterparts, resulting in an 81% and 5.7% increase in the adsorption capacity of Cd(II) and Pb(II), respectively.¹¹⁵

More importantly, the oxygen-containing functional groups on the weathered NPs can complex with HMs via inner-sphere complexes. For example, Huang et al. demonstrated that the formation of carbonyl and hydroxyl groups on weathered PLA and PS enhanced the complexation of Pb(II) and Ag(I).^83,116 Tang et al. revealed that amino N atoms within the amide group in polyamide 6 facilitated the adsorption of Cd(II), Pb(II), and Cd(II) via the formation of chemical coordination bonds, including Cu–N, Pb–N, and Cd–N.¹¹⁷ In addition, the hydrogen atom on the carboxyl group of weathered PS NPs (100–1000 nm) can form a hydrogen bond with the oxygen atom in the arsenic anion.¹¹⁸ Instead, a 33% decrease in the adsorption capacity of As(III) on PS NPs was observed when hydrogen bonds between them were broken at high temperatures.¹¹⁸ Furthermore, a significant increase in the ratio of aromatic C–H bands in light-weathered PS, coupled with an enhanced adsorption of Cd(II), suggested an important role of cation–π interaction in the adsorption of HMs on PS.¹¹⁹

The adsorption of HMs on weathered NPs can be influenced by the environmental factors, such as pH and NOM levels. The pH plays an important role in the adsorption of HMs by affecting the electronegativity of weathered plastics. For example, Holmes et al. demonstrated that PE had a negative charge at high pH, which increased the electrostatic attraction and thus increased the adsorption of Cd(II), Co(II), Ni(II) and Pb(II).¹²⁰ Natural organic matter can compete with weathered NPs for adsorption of HMs, thereby reducing the adsorptive interaction between NPs and HMs. Specifically, NOM reduced the adsorption of Cd(II) on polyamide, 2-propenitrile, and PET by 22%, 43%, and 45%, respectively.¹²¹ In addition, the formation of eco-corona by NOM can alter the hydrophobicity, polarity, and surface charge of NPs, thereby influencing the adsorption of HMs. Oxygen-containing functional groups of NOM-induced eco-corona increased the surface charges of NPs, thereby enhancing the electrostatic interaction with Pb(II).¹²² In addition, Guo et al. indicated that the formation of a polymer–HM–NOM ternary complex can enhance the adsorption of Cd(II) on PE, PP, PS, and PVC by 1.2-, 1.3-, 1.3-, and 1.3-fold, respectively.¹²³

3.2 Transformation of HMs by weathered NPs

Weathered NPs trigger the transformation of HMs, mainly due to the free radical-mediated redox processes (Fig. 4b). Chromium(VI) was reduced by light-weathered PS NPs (100, 1000 nm) via reductive O₂˙⁻.¹²⁴ However, long-term light weathering inhibited the reduction of Cr(VI) due to the increase in the amount of carbonyl groups on weathered PS NPs, resulting in the production of oxidative ·OH and ¹O₂ instead of O₂˙⁻.¹²⁴ Likewise, Ag(I) was reduced to Ag(0) in the presence of light-weathered PS NPs (100–1000 nm) via the formation of O₂˙⁻.¹³ Through the generation of ROO· and O₂˙⁻, the oxidation of Mn(II) was driven by light-weathered PS NPs (30–500 nm).¹²⁵ In addition, Ag(0) and ZnO were oxidized in the presence of light-weathered PS NPs (100 nm) via the generation of ·OH, ¹O₂, or acids.^34,126 The oxidation of metal sulfides can also be triggered by the presence of weathered NPs. For example, the oxidation of CdS and CuS was promoted by the light-weathered PS NPs (100 nm) through the production of ·OH and ¹O₂, resulting in the release of 30% and 2.5% of the total Cd and Cu, respectively.^34,126 While it is well documented that weathered NPs produce both reductive and oxidative free radicals in the laboratory, it is still unclear which free radical dominates in the field.

Notably, the observed Ag(I) reduction by light-weathered PS is not only due to the generation of free radicals, such as O₂˙⁻ production,³⁴ but is also facilitated by electron shuttling through carbonyl groups. For example, Ag(I) reduction was observed in the presence of light-weathered PS in freshwater and sand matrices.¹²⁷ The possibility of free radicals in Ag(I) reduction was excluded. Instead, a significant linear correlation was observed between the content of carbonyl groups on light-weathered PS and newly formed Ag(0).¹²⁷ The addition of Cu(II) (as an oxidizer of carbonyl groups) induced a concentration-dependent inhibition of Ag(I) reduction, suggesting a critical role of carbonyl groups on light-weathered PS in Ag(I) reduction.¹²⁷ These distinct mechanisms through either radical-mediated transformation or electron shuttling through carbonyl groups may be attributed to the chemical reactivity of PS induced before or after laboratory weathering. The former primarily results from polymer processing and/or storage.⁴² For example, we have identified the formation of free radicals on PS (1000 nm) prior to or after laboratory weathering (Fig. 3b).

In addition to the hitherto described ROS and carbonyl groups, other free radicals and functional groups derived from weathered NPs may provide new insights into their interaction with HMs. For example, the redox reactive SO₃˙⁻, with an oxidation potential of 0.7 V, can be produced from light-weathered PPS.¹²⁸ This species could influence the transformation of HMs, which deserves further investigation. In addition, recent studies have shown that after sulfur weathering (i.e., incubating with Na₂S solution), sulfur-containing functional groups were formed on the surface of plastics.¹²⁹ Whether sulfur-weathered NPs would drive the transformations of HMs, such as sulfidation, is largely unknown.

While the aforementioned NP-induced transformation of HMs is constrained in sunlit conditions, other environmental factors, including oxygen level, pH level, and cations, have been demonstrated to influence the processes. Oxygen, in particular, is often observed to facilitate these transformation processes directly and indirectly. Directly, oxygen serves as an oxidative agent, participating in the oxidative dissolution of HMs.¹³⁰ For example, oxygen can oxidize Ag(0) and CdS, releasing Ag(I) and Cd(II).³⁴ Indirectly, oxygen acts as a prerequisite for the formation of oxygen-containing groups or promotes ROS generation, which in turn indirectly influences the transformation of HMs. For example, the oxidative dissolutions of Ag(0) and CdS are strongly related to ROS produced from weathered NPs in an oxygen-rich atmosphere.^34,131 The pH level similarly influences NP-induced HM transformation, both directly and indirectly. For example, low pH conditions facilitate Ag(0) dissolution, whereas elevated pH conditions favor Ag(I) reduction. However, the reduction of Ag(I) induced by weathered PS NPs (100 nm) was hindered in low pH conditions.¹³² This is because O₂˙⁻ produced from weathered PS NPs can react with protons under acidic conditions to yield O₂·H, whose reductive ability is inferior to that of O₂˙⁻.¹³³ Cations, such as Ca(II), also exert an influence on NP-induced HM transformation via a reduction in the stability of NPs. Zhang et al. observed a decrease in the rate of Ag(I) reduction induced by PS NPs (100 nm) as Ca(II) was introduced.¹³² This is attributed to the promotion of (hetero-)aggregation of PS NPs and/or newly formed Ag(0) via the complexation of Ca(II) with oxygen-containing functional groups on weathered PS, which subsequently inhibited the ROS produced from PS.¹³²

More importantly, weathered NPs can fulfill the function of a driver of the transformation of HMs in a similarly effective manner to that of NOM. In the reduction processes, the observed Ag(I) reduction rates induced by weathered PS NPs (100 nm) ranged from 2.3 × 10⁻³ h⁻¹ to 6.8 × 10⁻² h⁻¹,^83,132 which are within the documented values for NOM (8.1 × 10⁻⁴–0.3 h⁻¹).^134–136 Similarly, the Cr(VI) reduction rate (2.3 × 10⁻³ h⁻¹) induced by weathered PS NPs (100–1000 nm)¹²⁴ is comparable with documented values for reduced-NOM (9.0 × 10⁻³ h⁻¹).¹³⁷ In the oxidation dissolution process, the CdS dissolution rates triggered by weathered PS NPs (100 nm) were 0.101–0.144 h⁻¹,¹³¹ aligning with the reported value ranges for NOM (3.9 × 10⁻²–0.199 h⁻¹).¹³⁸ Furthermore, Ag(0) dissolution rates induced by weathered PS NPs (100 nm) were documented to be 6.0 × 10⁻³–1.4 × 10⁻² h⁻¹,¹³² which were even more significant than that observed in certain natural water systems (5.0 × 10⁻³).¹³⁹ Collectively, these results suggest that weathered NPs are therefore significant, yet an unrecognized driver in nature to alter HM cycling.

4. Perspective

Despite the increasing number of studies on the risk assessment of NPs, their chemical reactivity has long been overlooked. This review identifies the chemical reactivity of NPs and clarifies their critical role in the transformation of HMs. However, many fundamental questions remain to be addressed, including those related to the weathering processes under environmental conditions, the quantitation of reactive NPs, and the risk of NPs in realistic environments.

4.1 Evaluating the weathering processes in realistic environments

While the chemical reactivity of NPs induced by weathering processes has been identified, the majority of investigations have been conducted under laboratory-simulated conditions. It is controversial whether the laboratory-simulated conditions represent environmentally relevant systems. For example, the light weathering of NPs in laboratory settings is conducted using a xenon lamp with a light intensity of 55 mW cm⁻², which is 1.6–3.3-fold greater than the daily solar irradiance received on Earth (16.4–34 mW cm⁻²).^34,140 The intensity and spectral distribution of sunlight are dependent on a number of factors, including latitude, longitude, solar elevation angle, and meteorological conditions.¹²⁰ Furthermore, NPs undergo a range of weathering processes that simultaneously occur in nature, which introduces complexity and uncertainty to the mechanisms involved. It is of great importance to gain a comprehensive understanding of the chemical reactivity of NPs by focusing on the weathering in large-scale field studies.

4.2 Estimating the pool of chemically reactive NPs in nature

The chemical relativity of weathered NPs can be quantitated by the levels of free radicals, functional groups, and released leachates, which can lead to an estimation of pool of chemically reactive NPs in nature. The unexpectedly large pool of reactive NPs may have significant consequences for the cycling of HMs as well as for carbon. For example, the formation of eco-corona on reactive NPs via adsorbing NOM could allow NPs to have a larger impact on carbon storage and transport.¹⁴¹ Based on the oxidation rate under light irradiation, we estimated a pool of reductive PS during light weathering in the natural environment (18.1–39.4 MT per day). However, this estimation does not account for the heterogeneity of NPs and their surroundings, where they experience varying degrees of weathering in distinct environmental matrixes, such as light weathering on the surface water but light-independent weathering in landfills and sediments. More work is needed to evaluate and differentiate the pool of reactive NPs in different environmental systems.

4.3 Focusing on the risk of NPs in nature

The observed complexity in the toxicity of NPs can be attributed to their chemical reactivity and intricate interaction with other chemical pollutants, such as HMs. A meta-analysis revealed that the combined exposure of micro(nano)plastics and HMs resulted in a significant reduction in root length, whereas exposure to micro(nano)plastics alone had no impact.¹⁴² This finding suggests that the toxicological effects observed in empirical studies focusing on NP particles may differ from those in realistic environmental conditions, where NPs are subjected to weathering and are highly likely to interact with HMs. It is therefore proposed that in order to obtain an unbiased result in the risk assessment of NPs, chemical reactivity can be regarded as a significant feature of NPs, along with their other properties, such as mass concentrations and size.

Data availability

Data availability is not applicable to this review article as no new data were created or analyzed in this study.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This work was supported by the National Natural Science Foundation of China (42207457), the National Key Research and Development Program of China (2023YFC3711400), and the Natural Science Foundation of Jiangsu Province (BK20220092).

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