key: cord-1044620-v4q6dmq2
authors: Liuwei, Wang; Wei-Min, Wu; Bolan, Nanthi S.; Tsang, Daniel C.W.; Li, Yang; Qin, Muhan; Hou, Deyi
title: Environmental fate, toxicity and risk management strategies of nanoplastics in the environment: Current status and future perspectives
date: 2020-07-08
journal: J Hazard Mater
DOI: 10.1016/j.jhazmat.2020.123415
sha: b19602f9a29082cc4a872a42e1a89e40ea46b15b
doc_id: 1044620
cord_uid: v4q6dmq2

Tiny plastic particles considered as emerging contaminants have attracted considerable interest in the last few years. Mechanical abrasion, photochemical oxidation and biological degradation of larger plastic debris result in the formation of microplastics (MPs, 1 μm to 5 mm) and nanoplastics (NPs, 1 nm to 1000 nm). Compared with MPs, the environmental fate, ecosystem toxicity and potential risks associated with NPs have so far been less explored. This review provides a state-of-the-art overview of current research on NPs with focus on currently less-investigated fields, such as the environmental fate in agroecosystems, migration in porous media, weathering, and toxic effects on plants. The co-transport of NPs with organic contaminants and heavy metals threaten human health and ecosystems. Furthermore, NPs may serve as a novel habitat for microbial colonization, and may act as carriers for pathogens (i.e., bacteria and viruses). An integrated framework is proposed to better understand the interrelationships between NPs, ecosystems and the human society. In order to fully understand the sources and sinks of NPs, more studies should focus on the total environment, including freshwater, ocean, groundwater, soil and air, and more attempts should be made to explore the aging and aggregation of NPs in environmentally relevant conditions. Considering the fact that naturally-weathered plastic debris may have distinct physicochemical characteristics, future studies should explore the environmental behavior of naturally-aged NPs rather than synthetic polystyrene nanobeads.

Made of various synthetic or semi-synthetic organic polymers, plastics are malleable materials capable of being molded into solid objects of various types and sizes. Due to the ease of manufacture, high stability and versatile properties, plastics have been used in a wide range of products. Thus, annual production of plastics keeps growing, reaching 359 million tons in 2018 [1] . Despite the fact that plastic recycling and management policies are improving, improper handling of plastic disposal is still a global trend, accounting for the unregulated release into the environment [1, 2] . Due to the hydrophobicity, physical and chemical resistance, plastics can be transported from terrestrial ecosystems to aquatic ecosystems. Plastics have been found in all kinds of environmental media, including the surface freshwater and the sediment, marine surface water and the seabed, groundwater, soil and even the atmosphere [3] [4] [5] [6] [7] [8] [9] [10] .

Once released into the environment, plastic particles are subjected to weathering and fragmentation (section 2.2). Various natural forces, such as the mechanical forces of water, ultraviolet (UV) radiation, and biological metabolism lead to the fragmentation into smaller plastic particles, namely microplastics (MPs) and nanoplastics (NPs). MP is defined as the plastic particles with the size ranging from 1 μm to 5 mm [11] [12] [13] .

Concerning NPs, there is still debate on its definition. Some scholars suggest that a J o u r n a l P r e -p r o o f definition of nanoparticles (from 1 nm to 100 nm) should be extrapolated to define NPs [14, 15] , while others adopted the whole nanometer range (from 1 nm to 1000 nm) [11, 16, 17] . In this review, we adopt the latter definition and regard sub-micron plastic particles (with diameter ranging from 100 nm to 1 μm) also as NPs.

Due to the recalcitrant characteristics of plastic particles, the environmental fate and the toxic effects of MPs have been widely explored. Many review articles have focused on various fields, including the sources [18] , distribution [19] , migration [20] , bioaccumulation [21] , toxicity [22] , ecological risks [23] , and remediation strategies [24] of MPs. In comparison, NPs are much less explored. Downsizing the plastic debris from micro to nano scale will result in a shift in physicochemical properties (section 2.2). Besides, the environmental behavior (such as aggregation and migration), bioaccumulation features, and toxicity of plastic particles are highly dependent on the size. Further investigations on NPs are necessary.

Similar with investigations on MPs, current studies regarding NPs have focused more on the marine ecosystem, especially the toxic effects of NPs on marine organisms, including bacteria, algae and fish [2, 25, 26] . Research on NPs in terrestrial and freshwater ecosystems is still limited. This review summarizes the current research size-based and density-based separation. As for size-based separation, the most widely used methods for nanoplastic enrichment are filtration (Figure 1 c) and field flow fractionation (FFF) (Figure 1 b) (Table 1 ). In (ultra)filtration systems, particles larger than the nano-sized membrane is collected (Figure 1 c) , and an external pressure could facilitate the water flow, thus increasing the operation speed. Field flow fractionation is a separation method where a field (e.g., electrical, centrifugal, gravitational, etc.) is applied perpendicularly to the fluid suspension crossing a long channel, resulting in the separation of particles present in the suspension depending on the differing mobilities under the external field-induced force [29, 30] . The most widely used FFF system is the Asymmetrical flow field flow fractionation (AF4), where the external field is the cross flow created by the asymmetrical wall (only the bottom wall of the channel is permeable) (Figure 1 b) . Due to the variance in diffusivity of particles (which is determined by size and particulate density), different particles are retained for different durations in the AF4 system (Figure 1 b) . The advantage of AF4 is that it can separate and characterize nanoparticles simultaneously through coupling to online detectors [11, 29] . Compared with size-based separation strategies, densitybased ones such as ultracentrifugation is seldom used in studies related to nanoplastics. This is probably because this technique has the limitation that it only processes small sample volumes (i.e., < 100 mL), limiting its applicability for environmental samples with low NP concentrations that require a much larger volume to obtain adequate amount of NPs [11] .

A number of techniques can be used to analyze physicochemical properties of NPs J o u r n a l P r e -p r o o f (Table 1) . Laser light scattering is the most widely used method for particle size assessment. When the laser passes through the suspension of NPs, a fluctuation of its intensity can be induced by the Brownian motion, which is dependent on the hydrodynamic diameter with NPs. Electron microscopy is widely adopted to investigate the surface morphology of NPs. Since the wavelength of electrons is much shorter than that of visible light, electron microscopies possess much higher resolutions (several nanometers) as compared with optical microscopies (> 1 μm). Besides, electron microscopies can be coupled with Energy-Dispersive X-ray Spectrometers (EDS) to investigate elemental distributions simultaneously. Although several conventional characterization methods such as Fourier-Transform Infrared Spectroscopy (FT-IR) and X-ray Photoelectron Spectroscopy (XPS) cannot be used to characterize a certain nanoplastic particle due to their size detection limits, they can provide much useful information for bulk samples of NPs [11] . For more detailed discussion regarding the separation and characterization methods of NPs, we refer readers to Schwaferts, Niessner, Elsner and Ivleva [11] and Fu, Min, Jiang, Li and Zhang [31] . Copyright 2018 Elsevier; (b) Separation of particles using asymmetrical flow field flow fractionation (AF4). Smaller particles possess higher diffusion coefficients, which stabilize further away from the membrane. Thereby, they are subjected to faster steamlines than larger ones, and exit the channel more quickly. Reproduced with permission from [33] . Copyright 2015 Frontiers Media; (c) Isolation of NPs from facial scrubs using five filtration steps. Reproduced with permission from [34] . Copyright 2017 American Chemical Society. 

The size of a plastic particle is the dominant characteristic determining its environmental fate (e.g., migration) [6, 52, 53] . Besides, bioaccumulation and toxicity J o u r n a l P r e -p r o o f can be size-dependent [54] [55] [56] . Considering that NPs mainly originate from the fragmentation and transformation of larger plastic particles (secondary NPs, section 3.1), investigating the downsizing mechanisms will be helpful for a better understanding of NPs.

NPs can be generated through the mechanical abrasion processes. The breakdown of daily-use polystyrene products by household blender generate considerable amounts of NPs [57] . Fragmentation of solid plastic wastes and MPs generate NPs in sewer system due the turbulence of water flow and mechanical devices in wastewater treatment plants (WWTPs) [58] . The natural fragmentation of larger plastic pieces can also be achieved in the sea swash zone [59] . (1)

where is the size of NPs, is the stress intensity factor of the plastic material, is the density, 0 represents the elastic wave speed, refers to the stain of plastic material, which is dependent on the applied stress.

Hydrolysis (react with water) is another potential mechanism accounting for NP

generation, yet it may not be the most powerful one at reducing the sizes of plastics [62] . In comparison, degradation initiated by UV irradiation is a very efficient J o u r n a l P r e -p r o o f downsizing mechanism. The photodegradation of plastics is mainly induced by reactive oxygen species. The decrease in particle size may be due to the chain scission by attacks from free radicals, such as hydroxyl (·OH), alkyl (R·), alkoxyl (RO·) and peroxyl (ROO·) radicals produced from the UV light. Possible reaction mechanisms for free-radical induced fragmentation include three steps (Eq. 2-9) [63] [64] [65] :

Step 1-initiation

Step 2-propagation Step 3-termination R • +R • → Not free radical products (7) R • +ROO • → Not free radical products (8) ROO • +ROO • → Not free radical products (9) Biological degradation and fragmentation of large plastic pieces and MPs by marine and terrestrial animals could also generate NPs in environment. The ingestion of plastic MPs and potentially NPs by marine organisms has been found among zooplankton, fish, shrimps and other animals [66] [67] [68] . Fragmentation or degradation of MPs into NPs has been reported in Antarctic krill (Euphausia superba) exposed to J o u r n a l P r e -p r o o f polyethylene MPs (31.5 μm) together with algal food. After ingestion, NPs of 150-500 nm size were formed, which were found in the digestive gland [69] . Reduction of MPs into smaller sizes has been observed in the common earthworm [70] and snails [6] , although fragmentation of MPs into NPs was not considered in these studies due to limitation of excess tools.

The morphology NPs are mainly determined by their origins (i.e., natural weathering vs synthetic fabrication). NPs from different origins have diverse shapes ( Figure 2 ).

Many studies regarding the migration, bioaccumulation and toxicity have adopted commercially-available NPs, which exhibit ideal spherical morphology in most cases Due to natural forces such as mechanical forces of water [71] , UV radiation [49] , and biological metabolism [72] , the shapes of resulting NPs become hardly smooth and spherical.

J o u r n a l P r e -p r o o f (d) synthetic metal-doped polyacrylonitrile (PAN) nanoparticle with a raspberry-like appearance [74] . All images are reproduced with permission.

Downsizing of plastic particles from micro to nano scale can also lead to a shift in chemical properties, especially surface functional groups. As shown in Eq. (2)-Eq. (9), reactive oxygen species are generated during the photodegradation process. This may result in an increase in oxygen-containing functional groups such as carboxyl, carbonyl J o u r n a l P r e -p r o o f and hydroxyl on the surface of NPs [63, 75] . The changes in surface functional groups alter the hydrophobicity and surface charges of NPs, which may affect the migration [76] , aggregation [77, 78] , contaminant adsorption [18] , bioavailability [79] and toxicity [80, 81] of NPs. It is therefore necessary to fully understand the weathering process of plastic particles. However, current studies mostly focus on the environmental behavior and toxic effects of synthetic spherical NPs since they can be easily obtained. It is argued that results from current studies may not reveal the behavior of naturally weathered NPs under field conditions (section 6).

are from land-based sources, such as coastal landfill operations, NPs carried by rivers and streams, biosolid and compost applications, and improper disposal of untreated sewage [2, 82] . Besides, direct marine-based sources include discharging of litters from ships/boats and fishing nets [17, 82] .

In this sense, understanding the terrestrial origin of NPs is crucial. One of the main sources of NPs is the domestic activities. Tiny fibers of polyester, nylon, acrylic and spandex are carried off to wastewater treatment plants during clothes laundry [83] .

Fragmentation of microbeads used in shampoos and scrubs release considerable amounts of NPs [34] . Even plastic tea bags could release billions of NPs [51] . Apart from domestic origins, industrial sources include the direct fabrication of NPs [84] and feedstocks of plastic products [17, 85] . In addition, agricultural activities contribute to the release of NPs. Application of sewage sludge as fertilizers represents a significant source [3, 7] , while plastic mulching [5] and polymer-coated slow release fertilizers and pesticides [18] present other potential origins of NPs.

Although migration characteristics of NPs in aquatic ecosystems are poorly investigated, methods and results from MPs may provide some insights in this field.

Considering the vastness of oceans, models have been developed to predict the migration of MPs in marine ecosystems. Both theoretical and empirical models indicate that the ocean currents redistribute the plastic particles in surface oceanic waters, J o u r n a l P r e -p r o o f which will accumulate in five major "garbage patches" located in North and South of Atlantic, North and South of Pacific, and Indian Oceans ( Figure S1 ) [86] [87] [88] [89] [90] . However, some scholars argued that an underestimation of plastics will occur if models focus merely on the surface ocean [87, 91, 92] . Instead, models should also take deep waters, coastal sediments, and deep-sea sediments into account. A comprehensive review of migration characteristics of microplastics in the ocean is provided elsewhere [87] .

For terrestrial ecosystems, current studies have mainly focused on the migration of NPs in porous media ( Table 2) . Although both artificial (e.g., quartz sand) and natural (e.g., natural sand, soil) solid materials can be used by these studies, the latter is preferred due to a more realistic size distribution. The transport of NPs is affected by several parameters, including particle size, ionic strength, surface functional groups of NPs, and organic matter (Table 2) . Song, Yang, Chen, Zhao, Zhao, Ruan, Wang and Yang [6] observed a trend of higher mobility of larger NPs (200 nm vs 50 nm) due to greater particle stability. An increase in ionic strength inhibits NP transport, since the compression of electrical double layer results in the formation of aggregates [93, 94] .

Dong, Zhu, Zhang, Huang, Lv, Jing, Yang, Wang and Qiu [76] found that retention of functionalized plastic particles in saturated sand followed the order of amino-> sulfonic-> carboxyl-modified NPs. They suggested that surface functional groups of NPs affect their affinity towards organic matter and ions in the aqueous phase, which may lead to different retention and migration rates. Carboxyl groups favored the adsorption of inorganic ions, hydrophilic contaminants, dissolved organic matter and J o u r n a l P r e -p r o o f suspended organic matter, while amino groups inhibited the adsorption of suspended organic matter and inorganic ions [6] . As shown in Table 2 , addition of humic acid, a fraction of natural organic matter, could significantly enhance the migration of NPs [76] .

This is because organic matter forms an eco-corona layer (coating) on NP surfaces, which prevents plastic particles from aggregation (section 3.3) and improves stability of NPs through enhanced steric and electrostatic repulsion [3, 45, 76] . However, not all types of organic matter will enhance migration. Song, Yang, Chen, Zhao, Zhao, Ruan, Wang and Yang [6] observed that suspended organic matter increased the stability and mobility of NPs, while dissolved organic matter decreased both. This was probably because that the coating of dissolved organic matter lowered the electrostatic repulsion between NPs and the sand, which allowed the attractive forces to be dominant, including Van der Waals force, organic polymer bridging, cation bridging, and electrostatic attraction (from positively-charged minerals).

On the one hand, co-existing contaminants/amendments affect the mobility of NPs.

Addition of biochar in quartz sand or soils resulted in decreased mobility of NPs, which was due to the formation of heteroaggregates involving biochar and NPs [53, 95] .

Sewage sludge application as a soil fertilizer may lead to elevated mobility of NPs owing to the release of dissolved organic matter [96] . Adsorption of naphthalene promoted the retention of NPs, which was attributed to the partial shielding of surface charge by the adsorbed nonpolar naphthalene molecule [94] . On the other hand, NPs may also affect the migration of heavy metals, organic molecules and even pathogens such as bacteria and viruses in porous media. Although no available data is available regarding the co-transport of NPs and heavy metals, it is inferred that weathered NPs may favor the transport of heavy metals in porous media, since the oxidized surfaces possess more oxygen-containing functional groups, favoring the surface complexation between NPs and metal cations [75] . Sorption of dissolved organic matter (DOM) onto NP surface also promotes heavy metal complexation (section 4.2.2), leading to the mobilization of heavy metals through NP-metal co-transport. NPs may interact with organic molecules through various ways, including heteroaggregation, hydrophobic interactions and electrostatic interactions. Dong, Zhang, Qiu, Yang, Wang and Zhang [97] have observed that polystyrene NPs act as a vehicle for the transport of fullerene (C60) through the decrease in ζ-potential (formation of NP-C60 heteroaggregates) (Table 2) . Liu, Ma, Zhu, Xia, Qi, Yao, Guo, Ji and Chen [98] have noticed that polystyrene NPs could enhance the migration of non-polar and weakly-polar molecules (e.g., pyrene, 2,2',4,4'-tetrabromodiphenyl ether) in soil, while revealing no significant effect on the transport of polar molecules (e.g., bisphenol A). This is because non-polar and weakly-polar molecules can be adsorbed in the inner matrices of polystyrene's glassy polymeric structure, resulting in physical entrapment of these organic contaminants. Co-transport of NPs and pathogens in soil threatens human health. With the outbreak of novel coronavirus (COVID-19/SARS-CoV-2), the presence of human pathogens in environmental compartments and their potential risks (e.g., groundwater

contamination as a result of soil-water transport, human inhalation due to soil-air transport) have raised much concern [99, 100] . Limited evidence has shown that the adsorption of NPs on bacteria facilitates the migration of E. coli through the repelling J o u r n a l P r e -p r o o f effect (Table 2 ) [52] . It is hypothesized that NPs may affect the mobility of bacteria and virus through various ways. Firstly, suspended NPs in the soil solution repel pathogens from approaching the soil colloid surface, thus increasing their mobility. In addition,

NPs may adsorb to the pathogen surface, forming a plastic coating that prevents the formation of large pathogen-pathogen aggregates. NPs may also compete with pathogens for binding sites on soil colloids directly, promoting the desorption and migration of bacteria and viruses. More studies on co-transport of NPs and pathogens are needed before drawing a clear conclusion how NPs in soil and sand affect the mobility of bacteria and viruses. 

The environmental fate of NPs is mainly governed by the weathering and the aggregation processes. Various stressors (environmental factors), such as the heat, water, UV irradiation, oxidants, microorganisms, or the combination of these causes the aging of NPs in the environment [101] . An elevation in temperature will accelerate the weathering of NPs as per the Arrhenius relationship [102] . The shear forces of the water cause the mechanical fragmentation (physical weathering) of NPs (section 2.2).

Artificial aging using UV and O3 co-exposure resulted in much rougher morphology and more oxygen-containing functional groups (e.g., hydroxyl, carbonyl, carboxyl) as compared with pristine NPs [75] . During this abiotic oxidation process, reactive oxygen species such as hydroxyl radical (·OH), singlet oxygen ( 1 O2) and superoxide radical (·O2 -) induced the chain reactions, which degraded the structure of NPs. Furthermore, oxygen was introduced to the surface of NPs, resulting in an increased number of oxygen-containing functional groups [75, 103, 104] . Microorganisms may also play vital roles in the biological weathering of NPs (section 2.2) through the colonization (plastisphere, section 4.2.3) and the utilization of the polymer matrix as a food source [101, 105, 106] .

Aggregation is a key issue in understanding the environmental fate of NPs. Evidence has shown that NPs can form milli-sized (mm-sized) aggregates in ecosystems [107] .

In addition, formation of heteroaggregates with inorganic colloids or organic matter lead to either settlement or migration of NPs [45, 97, 108] . In order to describe the aggregation process of NPs, the Derjaguin-Landau-Verwey-Overbeek (DLVO) theory [109, 110] has been widely adopted by various studies. The DLVO theory proposes that two independent forces, Van der Waals force (Eq. 11) and the electrostatic double layer force (Eq. 12) determine the stability of suspended particles [63, [111] [112] [113] : (12) where ( ) is the total interaction energy, ( ) is the Van der Waals interaction energy, and ( ) represents the electric double layer interaction energy in Eq. (10).

The parameters in Eq. (11) are defined as follows: A is the Hamaker constant for the NP dispersion system, whose value is dependent on the types of NPs and the aqueous media; a is the radius of NPs; d is the separation distance between NPs. As for the parameters in Eq. (12), is the dielectric constant of aqueous phase; 0 is the dielectric constant of vacuum; k is the Boltzman constant; T is the absolute temperature; q e is the electron charge; z is the charge number; is the surface potential of NPs (assumed to be equall to ζ-potential); is Debye length (Eq. 13):

where is the Avogadro constant, I is the ionic strength of the aqueous phase.

As shown in Eq. (12) , an elevation in absolute value of ζ-potential (| |) will lead to enhanced repulsive energy ( ( )) and total interaction energy ( ( )), making it NPs [93, 108, 111] . Natural minerals such as clay tends to form heteroaggregates with NPs due to electrostatic interactions [115] . Natural organic matter protects NPs from aggregation by elevating the | | value (due to the formation of eco-corona) [114, 116] .

However, Yu, Shen, Li, Fu, Zhang, Liu and Liu [77] suggested that if the concentration of organic matter is high enough (to enable the existence of un-adsorbed free organic matters in the system), the co-existence of natural organic matter and inorganic ions may lead to the "bridging effect". Inorganic metal cations (e.g., Ca 2+ ) could bridge oxygen-containing groups of both NPs and the organic matter, resulting in J o u r n a l P r e -p r o o f heteroaggregation.

The conventional DLVO theory can be adapted to describe NP interactions in more complicated systems (e.g., soil). Liu, Zhang, Tian, Liu, Qi, Ma, Ji and Chen [75] adopted an extended DLVO theory to assess the interaction energy of NPs in porous media. Total NPs-soil interaction energy takes three forces into account: Van interfering with chromatin structure and function. In this way NPs may induce genotoxic effects (cytogenetic anomalies and micronuclei) [119] . The internalization of NPs in various cellular compartments resulted in reduced root growth of onion [119] .

Interestingly, internalization of NPs may also have positive effects on plant growth.

After exposure to polystyrene NPs (0.01 -10 mg/L), root elongation of Triticum aestivum L. (wheat) was significantly (p < 0.01) enhanced by 89% -123%, as compared with the control [120] . Besides, increases in plant biomass, carbon and nitrogen contents were observed. NPs exposure resulted in enhanced growth of wheat seedlings without any overt stress. This was probably because polystyrene NPs increased the activity of α-amylase as a nanocatalyst, thus accelerating the production of soluble sugars from the starch granules [120] . However, NPs were also found to accumulate in wheat tissues (Figure 3 a, 

Ecotoxicity of MPs and NPs in the marine environment have been extensively investigated and reviewed [2, 26, 82, 121] . NPs can affect organisms from various trophic levels, including bacteria [122] , algae [123] , arthropods [124] , echinoderms [81] , J o u r n a l P r e -p r o o f bivalves [125] , rotifers [126] and fish [127] . Bioaccumulation of NPs in tissues [128] , effects of NPs on growth and reproduction [129] , NP-induced damages to immune system [130] , neurotoxicity [131] and alterations on metabolism pathways (especially the lipid metabolism) [132] are several major concerns in this field. A comprehensive review and detailed discussion on the toxic effects of NPs on marine organisms is provided elsewhere [2, 26] . pteridoides [135] . Due to environmental stress induced by NPs, an increase in male gametophytes were observed, which will cause drastic consequences for the reproductive success.

Inhalation, dermal exposure and ingestion are potential exposure pathways of NPs ( Figure 4 ). Inhalation of NP-containing aerosols and the penetration of NPs into the capillary blood system enable this nano-sized contaminant to distribute throughout the human body [136] . Evidence has shown that atmospheric fallout is a potential source of MPs and NPs [8] , but there is no available data concerning the concentration of airborne NPs. Wright and Kelly [137] suggested that sea salt aerosols, wind-driven Extrapolation of results from other nanoparticles may provide fresh insights into this topic. Evidence has shown that ingredients in skin care lotions (i.e., glycerol, urea and alpha hydroxyl acids) enhance the penetration of quantum dot nanoparticles (20.9 nm)

into excised human skin [140] . Although the surface chemistry of nanoplastics and other nanoparticles (e.g., quantum dot, metal oxide nanoparticles) may be different, the penetration of NPs is highly dependent on the particle size, so the results from nanoparticles may provide useful information on the penetration of NPs with similar sizes. In general, skin penetration may contribute to the human intake of NPs, but more direct evidences are needed.

Ingestion of contaminated food and water is probably the major exposure pathway of NPs. The gastrointestinal (GI) tract, with a large surface area of approximately 32 m 2 [141] , is the primary site for the uptake of NPs (Figure 4 ). NPs may cross intestinal villi and enter the blood vessel, and the formation of protein-plastic complex (so-called protein corona) is confirmed by in vitro studies [142] . This phenomenon is critical to the toxicity of NPs in organisms, since the interactions between tissues and organs occur with protein-coated, rather than bare NPs [138] . Results from an in vitro study of human blood cells indicate that protein-coated NPs can cause higher cytotoxic and genotoxic effect compared with virgin NPs [142] . This is probably because the formation of biomolecular corona on the surface of NPs helps them escape from the immune system, resulting in prolonged existence in the circulation system. The binding mechanism of protein corona with NP is not well understood, but it is believed that non-J o u r n a l P r e -p r o o f 31 specific physical attraction (i.e., Van der Waals force), hydrogen bond, and delocalized π bond contribute to the formation of this protein-plastic complex [142, 143] .

A limited number of rodent in vivo and human in vitro studies have shown that NPs can have adverse effects on the immune system. Toxic effects of NPs on human cells include induced up-regulation of cytokines involved in gastric pathologies [144] , disruption of iron transport [145] , induction of apoptosis [146] , endoplasmic reticulum stress [147] and oxidative stress [148] . One feasible way for the better understanding of NPs' toxicology is the extrapolations from nanoparticles. However, Bouwmeester, Hollman and Peters [149] suggested that the extrapolation of information on nanoparticles to nanoplastics should be conducted with care, since NPs possess evidently different surface chemistry. For further information regarding the toxicology and the feasibility of extrapolation, we refer readers to Lehner, Weder, Petri-Fink and Rothen-Rutishauser [138] and Bouwmeester, Hollman and Peters [149] . 

Due to the large specific surface area of NPs, various contaminants such as polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), and heavy metals can be sorbed on their surfaces [18] . The ability of contaminant sorption on NPs can be described using a sorption coefficient [150] (Eq. 14):

where is the sorption coefficient, * and * refer to the concentrations of contaminants in plastic and aqueous media, respectively (at equilibrium). The sorption coefficient is crucial for the understanding of sorption and desorption behaviors of contaminants. If the actual Cp/Cw ratio is below Kpw, adsorption occurs. If the Cp/Cw J o u r n a l P r e -p r o o f ratio is above Kpw, desorption takes place. If the Cp/Cw ratio falls equal to Kpw, the adsorption-desorption system reaches equilibrium. Equilibrium can occur in natural conditions [150] . For instance, in marine ecosystems, sorption of contaminants to floating NPs may proceed till equilibrium, since the contaminant concentration is relatively stable over time. However, in a wastewater treatment facility, sorption equilibrium is less likely to occur, since contaminant concentrations fluctuate with time.

Understanding the sorption behavior of NPs is quite crucial in assessing the toxicity of various contaminants. Organisms accumulate contaminants via various pathways ( Figure 5 ), including the direct uptake of free available contaminants and the uptake of NP-adsorbed contaminants. It is noteworthy that contaminants can either desorb from or still attach to NPs after bioaccumulation [151, 152] . Results from toxicological research on MPs [3, 153] and nanoparticles [154, 155] can be extrapolated to propose several potential mechanisms. Firstly, higher specific surface areas of aged NPs favor the contaminant adsorption, which may decrease the biological uptake of desorbed contaminants. Secondly, the increase in polarity of NPs (e.g., introduction of carbonyl groups) during photo-oxidation may elevate the risks of NP-associated non-polar organic contaminants, while decreasing the risks of heavy metals (as a result of enhanced surface complexation). Furthermore, the hydrophobicity of contaminant itself affects the toxicity, since hydrophilic contaminants are more likely to be desorbed from NPs after uptake. As has been discussed above, the sorption coefficient plays a vital role in adsorption-desorption of nanoparticles. However, this parameter is often 

Discrepancy exists whether NPs can make organic contaminants more toxic (Table 3) .

Several studies have found that toxicity can be greatly enhanced due to the high adsorption capacity towards hydrophobic organic contaminants such as polychlorinated biphenyls (PCBs) [156] and bisphenol A [152] . Enhanced adsorption may result in enhanced bioaccumulation. For instance, Chen, Yin, Jia, Schiwy, Legradi,

Yang and Hollert [152] found that polystyrene NPs enhanced the bioaccumulation of bisphenol A in head and viscera of zebrafish, which may stem from the enhanced uptake of NP-adsorbed form. Ma Several reasons may account for this discrepancy ( Figure 5 ). Concentration of NPs affects the sorption process greatly. For instance, when NP concentration is very low (i.e., 0.4 mg/L), NP-adsorbed pyrene accounted for less than 1% of the total pyrene bioaccumulation in Perinereis aibuhitensis [151] . Pyrene is mostly free available in this case, so NPs had little effect on its toxicity. Besides, the particle size of NPs affects the number of available sorption sites, thus affecting the sorption coefficient. For instance,

Velzeboer, Kwadijk and Koelmans [160] found that NPs enhanced PCBs sorption by 1-2 orders of magnitude as compared with MPs. In addition, toxicity of certain contaminants is species-dependent [161, 162] , so the selection of different species may result in varied results. Variance in hydrophobicity/polarity of organic contaminants also results in different sorption capacities. Hydrophobic organic contaminants tend to adsorb onto NPs (Kpw > 1), indicating that an enrichment in plastic J o u r n a l P r e -p r o o f phase rather than aqueous phase will occur. Polarity affects the adsorption mechanisms. Polar compounds bind with NPs through surface adsorption, while nonpolar/weakly polar compounds tend to adsorb in the inner matrices of NPs (physical entrapment) [98] . This may lead to the variance in bioavailability and mobility.

As shown in Table 3 , most of the studies used organisms living in aquatic ecosystems (e.g., zebrafish, Daphnia magna, clamworm). To date, the toxic effects of NP-attached contaminants on terrestrial ecosystems have not been fully investigated. Limited evidence have shown that NP-attached contaminants affects soil microbiome [163] .

Several other challenges and future directions in this field are discussed in section 6. Note:

*Organisms in terrestrial ecosystems J o u r n a l P r e -p r o o f

Unlike organic contaminants, heavy metals cannot be degraded, rendering wide distribution due to both natural and anthropogenic sources [166, 167] . Bioaccumulation of toxic metals along the food chain poses a severe threat to human health [168] [169] [170] [171] .

However, the toxicity of NP-adsorbed heavy metals is poorly investigated. As has been Pristine particulate plastics with high hydrophobicity have a lower likelihood of interacting with heavy metals when compared to particulate plastic-DOM assemblages.

In the latter case, there is greater interaction and increased retention of trace elements.

For example, Wijesekara, Bolan, Bradney, Obadamudalige, Seshadri, Kunhikrishnan, Dharmarajan, Ok, Rinklebe, Kirkham and Vithanage [173] identified the adsorption of heavy metals (i.e., Cu) onto particulate plastics that had modified surfaces due to DOM J o u r n a l P r e -p r o o f adsorption. The findings implied that modified particulate plastics adsorbed significantly greater concentrations of Cu than pristine particulate plastics, possibly due to the introduction of oxygen-containing functional groups (enhance surface complexation). Furthermore, long term pre-modification (e.g., photooxidation and attrition of charged materials) that contributes to aging of plastics causes these aged particles to have a great metal sorption capacity.

Previous studies indicate that many bacteria can attach/grow on various plastics surface in either aquatic environment or ambient conditions. The concept of "plastisphere" was proposed based on observation that plastics function as habitats and are rapidly colonized by marine microorganisms [105, 174, 175] . The plastisphere is the layer of microbial life that forms around every piece of floating plastic. Plastics may not only serve as a novel microhabitat for biofilm colonization, but also increase the likelihood of pathogens propagating. Evidence has shown that floating plastics in aquatic ecosystems can act as vectors (a means of transport) for pathogens such as

Vibrio [174] and Pseudomonas [176] . A study by Huang, Zhao, Wang, Zhang, Jia and Qin [177] even observed the enrichment of potential human pathogens Nocardiaceae,

Campylobacteraceae and Vibrio in soil as a result of LDPE film application. Apart from the selective enrichment of pathogens, the abundance of antibiotic resistance gene (ARG) may also increase within the plastisphere. Wu, Pan, Li, Li, Bartlam and Wang [176] found that the ARG abundance of the plastic biofilm was 3-fold higher than that of the river water samples, indicating the enrichment of ARGs by the plastisphere.

Wang, Xue, Li, Zhang, Pan and Luo [178] noticed that the adsorption of antibiotics lead to a significant shift in ARGs (i.e., sul1, tetX and ermE) on PE plastics. In addition, plastics in the soil may also increase the abundance and act as a sink of ARGs, affecting soil health in the long term [179, 180] .

Recently, a study on COVID-19 on contaminated surfaces and in aerosols suggests that people may acquire the coronavirus through the air and after touching contaminated objects. The virus was detectable for up to three hours in aerosols, up to four hours on copper, up to 24 hours on cardboard and up to two to three days on plastic and stainless steel [181] . This has raised a question about potential risk of airborne disease by coronavirus and flu virus via plastic MPs and NPs in polluted air.

It does take scientists decades to develop the concept of NPs inevitably, because there's adequate evidence indicating that plastic substances do break into NPs and remain well detectable all around the environment [182] , and reluctantly, because their more widespread distribution and harder isolation methods compared with MPs [34, 183] will have led to somehow severer threats on living creatures, human society and even the whole ecosystem. To better understand the risk associated with NPs, a DPSIR (driving forcespressuresstatesimpactsresponses) framework can be adopted ( Figure 6 ). Due to the increase in population and economic growth, global 

Due to the substantial gap of knowledge regarding NPs (e.g., their amount in the environment, environmental behaviors, exposure pathways), challenges remain in the risk assessment of NPs. To better understand the risks associated with NPs, substantial scientific efforts are imminent. The first step, however, is the acquisition of robust data on exposure in marine, freshwater and terrestrial settings [184] . Currently they are very limited due to the lack of precise analytical approaches (section 2.1), so the environmental concentrations of NPs are only estimates [11, 184] . For further information regarding the risk assessment of NPs, readers are referred to Alexy, Anklam, Emans, Furfari, Galgani, Hanke, Koelmans, Pant, Saveyn and Sokull

Kluettgen [185] and Pinto da Costa, Reis, Paço, Costa, Duarte and Rocha-Santos [38] , who have provided comprehensive overviews on this topic.

Solutions to NP pollution basically lie in the following aspects: 1) development of novel remediation technologies, 2) policy making and 3) public awareness (Figure 7) .

Although little effort has been made regarding the remediation strategies of NPs, several possible directions have been proposed by present studies. Biotechnology J o u r n a l P r e -p r o o f advances make it possible to gradually replace conventional non-biodegradable plastic products by biodegradable ones [186] , while special enzymes, bacteria and fungi that are capable of degrading plastics can be introduced to promote the disposal process [187] . The biodegradation process can be divided into three stages [188, 189] :

Stage 1: depolymerization of plastics into monomers or oligomers. This process, which is induced by extracellular enzymes, takes place outside the microorganisms.

Stage 2: assimilation of monomers and oligomers. The depolymerized products enter the cell, and become part of the living biomass of microorganisms through metabolism.

Stage 3: mineralization. The assimilated plastic metabolites will be oxidized, forming CO2 and H2O.

To prevent NP from entering the aquatic ecosystems, the most effective engineering technology is to remove NPs in a wastewater treatment facility. Pre-treatment strategies including density separation and coagulation as well as membrane separation have proven effective for the removal of MPs in drinking water, but their ability to remove NPs should be further examined [190] . As has been discussed in suggests that addition of soil amendments (e.g., biochar) could reduce the migration of NPs in porous media, thus mitigating the risks through retention/stabilization [53] . It J o u r n a l P r e -p r o o f is herein proposed that remediation strategies towards other contaminants can be further extrapolated to NPs. For instance, addition of "green" remediation materials, such as engineered biochar [191, 192] and clay minerals [193, 194] , have proven effective for the immobilization organic contaminants in soil. Considering the organic polymeric nature of NPs, the mechanisms involved in organic contaminant immobilization and NPs retention may be quite similar (e.g., π-π interactions, hydrogen bonding). Therefore, successful attempts in organic contaminant stabilization may shed light on the remediation of NPs in terrestrial systems (especially the soil).

Policy making is the most efficient and reliable way to keep NPs-related risks under

control, yet one should always keep in mind that a policy can only be promulgated after its effects have been carefully assessed [195] . In practice, evidence-based policies are able to tackle the problems caused by NPs in the whole life cycle, such as the US Microbead-Free Waters Act (acting on the source) [196] , the Packaging and Packaging

Waste Directive (94/62/EC) (acting on the using stage) [197] , and the Directive on the Landfill of Waste (1999/31/EC) (acting on the disposal stage) [198] . Regarding the fact that attention to emerging contaminants by policy makers often peak a few years later than scientific attention [199] , more interactions between policy makers and researchers are encouraged to bring useful findings into practice as soon as possible.

Since plastic products are an essential part in daily life, raising public awareness of the NPs could be a feasible and advantageous solution to manage the potential risks of NPs. As could be expected, this particular aspect is fundamental but inefficient, as it usually takes months or years to alter people's thoughts and attitudes, not to mention J o u r n a l P r e -p r o o f that such process ought to be supported by concrete scientific evidence [200] . 

Although research on marine plastics remains at the forefront, researchers have begun to address the concern of NPs in other ecosystems. The largest gap in current research is the understanding of the environmental behavior and ecologic impacts of NPs in terrestrial systems. Due to the lack of available information regarding the concentrations, migration characteristics, bioaccumulation risks and toxic effects of NP particles, it is hard to assess whether these tiny particles will affect the well-being of terrestrial ecosystems. As discussed in section 3.2, migration of NPs in porous media (e.g., soil, sediment and sand) have been investigated by recent studies. However, it is argued that most of these studies have neglected the role of natural forces (such as rainfall, freeze-thaw, etc.) on NP migration, with only one exception that considered the rainfall process [96] . Besides, NPs may even enter the groundwater through vertical J o u r n a l P r e -p r o o f colloidal migration, which has been neglected by most studies [201] . Although several attempts have been made to explore the bioaccumulation of polystyrene NPs by microorganisms and plants, controversy still exists whether NP particles can induce toxic effects on organisms living in terrestrial systems. Some studies even observed that the presence of NPs can enhance the plant growth, but the mechanisms involved in this process remain unknown [120] . Future explorations on the effects of NPs on the agroecosystem should be given higher priority, since it is very likely that soil may act as a long-term sink for NPs and that NPs in crop tissues may be ingested, threatening the food safety in the long run [5, 202] .

Although information from marine plastic studies can be extrapolated to freshwater to some extent, researchers should bear in mind that NPs in freshwater will cause threats to the environment in their unique ways. For instance, apart from toxic effects on microorganisms and fish, NPs may affect the growth and reproduction of aquatic plant species as well [135] . In freshwater systems, a shift in ionic strength and dissolved organic matter content will result in distinct aggregation behaviors of NPs, which may affect the sedimentation process greatly. In addition, given the lack of terrestrial studies to date, current knowledge on the freshwater systems, especially the migration characteristics of NPs in the sediment and the phytotoxicity of NPs, may be helpful to infer the environmental behavior and toxicology of NPs in the terrestrial ecosystems.

It is widely acknowledged that NPs are widely spread in the hydrosphere, but research on the presence, transport and toxicity of NPs in the atmosphere are scarce. Evidence is mounting that inhalation of airborne particles containing NPs can induce toxic effects J o u r n a l P r e -p r o o f on human lung cells [136, 138, 203, 204] , but the origins and the concentrations of NPs in the air remain unknown. MPs have been detected in the atmosphere of urban [205, 206] and even remote mountain areas as a result of long-range transport [207, 208] . Moreover, airborne plastic particles will enter terrestrial and aquatic systems through deposition [209] [210] [211] . The atmosphere may also serve as a "superhighway" for NPs. Future research on the airborne NPs is desperately needed.

According to the literature reviewed, the most frequently used nanoplastic particle is the polystyrene (PS). However, this is not consistent with the global plastic demand. It is estimated that the polyethylene (PE), polypropylene (PP) and polyvinyl chloride (PVC) hold significantly higher proportion of annual plastic production (29.7%, 19.3%

and 10%, respectively) as compared with polystyrene (PS), which only accounts for 6% [1] . Considering the fact that various types of NPs possess distinct physicochemical properties (i.e., functional groups, polarity, hydrophobicity, etc.), it is doubtful whether these studies can reveal the environmental fate and toxicology of nanoplastics from a wide range of polymers used for plastic manufacture. For instance, plastic mulching is a widely-adopted technique to promote agricultural production in many countries [5, 212] . It is estimated that plastic mulch film covers over 20 million hectares of farmland in China, which is a dominant source of NPs in the agroecosystem [213] . However, plastic mulch films are mainly made of PE rather than PS. When it comes to the migration and bioaccumulation characteristics of plastic particles in the farmland soil, extrapolating the results from PS NPs is somehow inappropriate, and can be J o u r n a l P r e -p r o o f misleading to some extent. Besides, PE mulch films usually contain considerable amounts of phthalate esters (PAEs) that can be released into soil. Current studies have failed to assess the toxic effects of these additives.

As has been discussed above, commercially-available synthetic polystyrene spheres may not reveal the real biogeochemical processes of NPs derived from plastic debris.

Once released into the environment, plastic debris undergoes the weathering process.

Mechanical forces (e.g., wave and current), biodegradation and photodegradation result in the fragmentation of plastic particles and a shift in physicochemical properties.

Only a few studies have investigated the aging process of NPs, and used naturally High concentrations of NPs (i.e., mg/L) have been adopted by most studies to assess the environmental fate and toxic effects of NPs. However, these concentrations could be hypothetically several magnitudes higher than environmental estimates [121] .

Currently the distribution and concentrations of NPs in marine, freshwater and soil environments is largely unknown due to restrictions on limits of detection (LOD). If assumed as ideal spheres, the mass of a NP particle will decrease with the third power of its diameter. Although the environmental sample of NPs can contain high particle numbers, low masses may hinder the analysis of concentrations [11] . More advanced technique with lower LOD is therefore required for the determination of nanoplastic concentrations in the environment.

The "unrealistic" also refers to the co-existing materials such as particulate and dissolved organic matter. Several studies examined the migration of NPs in pure watersaturated porous media [75, 96] . Other studies overlooked the roles of natural minerals, dissolved organic matter and suspended organic matter on the migration, aggregation, and toxic effects of NPs. Besides, in natural conditions, co-existing heavy metals and organic contaminants may adsorb onto NPs, affecting their transport and ecotoxicity.

Microorganisms attached onto NPs may even form biofilms on the NPs [214] [215] [216] . It is indeed impossible to take everything into account in a certain study, but researchers should spare no effort to predict the environmental behaviors and toxicology of NPs in a more realistic way. Compared with aquatic ecosystems, the environmental behaviors of NPs in terrestrial ecosystems are much more complex due to the heterogeneity of the environmental media (e.g., soil) and intensive anthropogenic activities. Since such J o u r n a l P r e -p r o o f a complex system cannot be duplicated in the laboratories, more field studies are therefore recommended to enhance our understanding towards the environmental fate, toxic effects and potential risks of NPs.

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

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