Authors: Sheng Wang, Noor Sulaiman AL-Hasni, Zhaoli Liu, Airong Liu
Categories: Aquatic environmental behavior, Differences, Microplastics, Nanoplastics, Review
Source: Environment & Health
Environmental Differences between Nanoplastics and Microplastics: Behavior and Fate
Microplastics and nanoplastics are emerging pollutants of concern in the aquatic environment that are causing increasing global environmental and human health problems. Although there has been extensive research on microplastics and nanoplastics, little has been said about the differences in their behavior in the aquatic environment, and many studies have considered them as the same class of hazardous materials; but in fact, microplastics and nanoplastics should be considered as two different types of environmentally hazardous materials. In this review, we propose that microplastics and nanoplastics behave in the aquatic environment in a size-dependent manner and should be distinguished. And we systematically analyzed the differences in the behavior of microplastics and nanoplastics in the aquatic environment in terms of five 1) distribution behavior; 2) adsorption behavior; 3) reaction with natural colloids; 4) aging and leaching behavior; 5) interaction with organisms. This paper has been written to draw academic attention to the different behaviors of microplastics and nanoplastics in the aquatic environment in order to distinguish between their effects on humans and the environment.
Keywords: Microplastics, Nanoplastics, Aquatic environmental behavior, Differences
In recent decades, the production of plastics has increased dramatically, with 359 million tons of plastics produced annually, and a cumulative total of nearly 10 billion tons of plastics have been produced worldwide.^1,2^ Inevitably, these plastics end up in the environment, and due to the high stability and corrosion resistance of plastics,^3^ it takes hundreds of years for them to be fully degraded when they enter water bodies and soils.^4^ Plastics have been found from mountain ranges to the seabed, from the Arctic to the Antarctic, and some researches have even shown hundreds of thousands of nanoparticles of plastic per liter of bottled water, posing serious health risks to human drinking water, making us concerned about the environmental behavior of plastics.^5,6^
The aquatic environment is one of the most important sinks for plastics.^7,8^ With approximately 5.25 trillion plastic fragments floating in the oceans and more than 2.36 tonnes of dissolved organic carbon (DOC) leaching from marine plastics globally each year, this DOC will have an impact on marine microbial food web and the carbon cycle, and this number will continue to grow in the future, with potentially unaccounted consequences for marine microbes and for the ocean system.^9^ In the environment, plastics are undergo physical, chemical, and biological changes, and large pieces of plastics are broken down into small plastic particles of different sizes, of which those with a diameter range of 1 μm-5 mm are called microplastics (MPs),^10,11^ and those with a diameter range of 0–1 μm are called nanoplastics (NPs).^12,13^ MPs and NPs are also collectively referred to as micro- and nanoplastics (MNPs), but due to their different particle sizes, they tend to exhibit different environmental behaviors,^4^ and therefore environmental behavior and fate are also important in differentiating between MNPs. As a new type of pollutant, MNPs have been spread globally due to their characteristics such as small size and high mobility, causing major environmental impacts.^14^ For example, 26 types of MPs have been detected in the soil of the southern Tibetan Plateau in China with an abundance of 4.01 pcs/g.^15^ MPs abundances of 200–2200 pcs/L have been detected in sediments from the deepest part of the Earth, the Mariana Trench.^16^ MPs abundance at some detection sites in the Arctic was 42–6595 pcs/kg.^17^ MPs were also detected in Antarctic snow samples at an abundance of 29 pcs/L.^18^ Recent reports indicate that MPs have been found in human blood,^19^ and similarly reports have shown that maternal exposure to polystyrene nanoplastics (PSNPs) increases the risk of neurodevelopmental defects in the offspring, leading to brain abnormalities in the offspring.^20^ The environmental impacts of plastic particles are widespread and plastics can cause serious damage to ecosystems, with a large body of literature citing a variety of deleterious effects of plastics on ecosystems.^21,22^ For example, MPs have affected aquatic organisms along the Caribbean coast of Colombia, leading to an increase in the activity of disease-causing microorganisms.^23^ These reports remind us that it is urgent to study the environmental behavior of MNPs.
However, the existing literature tends to consider NPs as smaller MPs or extensions of MPs, and indiscriminately explores the behavior of both in relation to the aquatic environment, with little attention paid to the differences in the behavior of NPs and MPs in the aquatic environment and the reasons for these differences. The limited literature that focuses on the environmental differences of MNPs tends to focus on a very small aspect, such as the differences in aggregation and sedimentation of MNPs,^24^ and the differences in the effects of MNPs on the gut microbial community,^25,26^ etc. There are fewer systematic descriptions of the differences in the environmental fate of MNPs among themselves. This is not conducive for researchers to understand the differences between NPs and MPs in the aquatic environment. Therefore, this review systematically elucidates the differences in environmental behavior between MPs and NPs in the aquatic environmental media in various aspects, and proposes that NPs are a special class of pollutants different from MPs, which will help researchers to comprehensively understand the aquatic environmental behavior of MPs and NPs.
According to global annual pure polymers (resins) production data from 1950 to 2015 published by the Europe Plastics Market Research Group, and global annual fiber production data from 1970 to 2015 published by The Fiber Year and Tecnon OrbiChem, a total of approximately 6.3 billion tonnes of plastic waste was generated from 1950 to 2015, which enters the environment in a variety of forms and may decompose into NPs.^27^ Since MPs are more easily separated than NPs during water treatment, and since individual MPs eventually break up into hundreds of NPs under continuous environmental exposure, for example in photoaged simulated seawater, MPs with a mass ratio of 7.03 ± 0.37% break up into NPs,^28^ therefore, aquatic organisms are more likely to be exposed to NPs than to MPs as the particle concentration of NPs will eventually significantly exceed that of MPs. The mass concentrations of microplastics and nanoplastics in the influent water of a wastewater treatment plant (WWTP) were found to be in the same order of magnitude, 26.23 μg/L and 11.28 μg/L, respectively, which implies that the number of particles of nanoplastics far exceeds that of microplastics,^29^ and the concentrations of microplastics and nanoplastics in other common aquatic environments are listed in Table 1. However, ecotoxicological studies of MPs and NPs have focused on the laboratory scale, and these studies often use contaminant concentrations that are many times higher than those found in actual water bodies and are of little relevance for reference in the actual aquatic environment.^30,31^ Therefore, future research is needed to further analyze the environmental risks of NPs and MPs in the real aquatic environment.
The motility of NPs and MPs in the aquatic environment is characterized by two main aspects, namely, the motility of plastic particles with water and their Brownian motion due to the impact of water molecules. The motility of plastic particles with water is mainly in the form of lateral transport and longitudinal settling or floating. In terms of lateral motion, compared to MPs of the same material composition, NPs are more easily carried and transported by water currents due to their smaller size, making them more susceptible to perturbation by water currents, whereas MPs are more easily retained due to their larger size.^34,35^ In terms of longitudinal motility, NPs have very little longitudinal movement compared to MPs due to their colloidal nature.^36^ Settling or floating is controlled by the density of the plastic particles, for example, low density polyethylene (PE) and polypropylene (PP) are both less dense than water, so it is expected that NPs and MPs that are less dense than water will eventually float to the surface of the aquatic environment, whereas materials such as high-density polyethylene, polyvinyl chloride (PVC), and polyurethane (PU) are denser than water, and therefore settling of these types of plastic is expected to occur.^37,38^ The settling velocity of plastic particles is affected by the particle size, and according to the Stokes’ settling equation, a rough calculation shows that the settling velocity of polyurethane microplastics (PUMPs) with a density of 1.38 g/cm^3^ and a diameter of 200 μm in water is about 8.275 × 10^–3^ m/s, whereas the settling velocity of polyurethane nanoplastics (PUNPs) with a diameter of 200 nm in water is only 8.275 × 10^–9^ m/s.^39^ More intuitively, the amount of PUNPs with a diameter of 200 nm that settles in one year (about 0.26 m) is even smaller than the amount of MPs of the same material with a diameter of 200 μm that settles in 1 min (about 0.5 m), showing that the longitudinal mobility of NPs in the aquatic environment is far inferior to that of MPs.
According to the principle of Brownian motion, each particle in the water has its own random movement of Brownian motion phenomenon. The smaller the particle size the fewer water molecules are hit from different directions at the same moment. According to the statistical law, when a small number of water molecules act simultaneously on the surface of particles, the combined force exerted on the particles cannot be balanced, and due to the fact that the mass of the small particles is much smaller, this results in the large acceleration of the small particles subjected to the combined force, so that the Brownian motion of NPs made of the same material is more obvious in the aquatic environment as compared to that of MPs.^40,41^ However, the Brownian motion does not affect the lateral and longitudinal movement of the NPs because the Brownian motion is irregular and unpredictable, and its speed and direction are random, so it does not cause macroscopic particle displacement.
of Adsorption Behavior
The adsorption behavior of plastic particles in the aquatic environment is not negligible, where MPs and NPs interact with particles and solutes in the aquatic environment through mechanisms such as electrostatic interactions, van der Waals forces, hydration, and chemisorption.^42,43^ Plastic particles in the aquatic environment play the dual role of adsorbate and adsorbent, so their adsorption behavior in the aquatic environment should be divided into two aspects. On the one hand, the plastic particles are able to adsorb smaller substances in the water, the most common being the adsorption of plastic particles on organic matter and heavy metals; on the other hand, plastic particles can be adsorbed on suspended substances and sediments in the water, such as silt and minerals, which affects the distribution of plastic particles.^44,45^
The adsorption behavior
of plastic particles on other substances in the aquatic environment
is complex and is influenced by a number of factors, one of which
is the particle size of the plastic particles.^46^ Experimental studies have shown that the adsorption behavior
of MPs and NPs differ significantly in the aquatic environment. Figure 1 summarizes the adsorption
behavior of MPs and NPs on different substances in the aquatic environment,
the specific data are presented in Table S1. As shown in Figure 1 a, the adsorption capacity of NPs (blue background) is usually greater
than that of MPs (pink background), which may be related to their
larger specific surface area and more active sites.^47^ Some studies have shown that the adsorption capacity of
PS plastic particles for As(III) and As(V) increases with decreasing
particle size. This is because the higher the point of zero charge
(PZC) of PS plastic particles as the particle size decreases, the
stronger the electrostatic adsorption capacity of arsenic will be,
and at the same time, due to the increase of the specific surface
area, the adsorption sites for arsenic also increase.^47−49^ The results of the adsorption of triclosan by PS plastic particles
of different particle sizes showed that PS particles of 41 μm
adsorbed almost no triclosan, while PS particles of 78 nm adsorbed
all the triclosan within 1 h.^50^ However,
due to the fact that NPs are very easily aggregated, this also contributes
to some extent to the fact that the adsorption capacity of NPs may
be lower than that of MPs. For polystyrene microplastics (PSMPs),
shown in Figure 1 b,
the Kd values of both phenanthrene and
nitrobenzene increased with decreasing particle size (pink background),
however, for PSNPs, the Kd values of phenanthrene
and nitrobenzene were significantly lower when the particle size decrease
(blue background), which was attributed to the fact that the aggregation
of the NPs considerably reduced the effective surface area available
for adsorption.^51^ Experiments have shown
that PVC particles smaller than 200 nm in size are easily aggregated,
resulting in a corresponding decrease in effective adsorption sites.^52^ Therefore, when comparing NPs and MPs as adsorbents
for substances undergoing adsorption in the aquatic environmental
media, it is essential to take the aggregation of the NPs themselves
into account. Influenced by the type of adsorbent, NPs cannot simply
be considered as MPs with smaller size and larger specific surface
area.
Figure 1 Adsorption behavior of MPs and NPs on different substances in the aquatic environment. (a) Adsorption capacity of MPs and NPs for As, the data in the left and right graphs were taken from different experiments. Reproduced with permission from ref (47). Copyright 2019, Elsevier, and reproduced with permission from ref (48). Copyright 2022, Elsevier. (b). Distribution coefficient of MPs and NPs for nitrobenzene (left) and phenanthrene (right). Reproduced with permission from ref (51). Copyright 2019, Elsevier.
When plastic particles are acting as adsorbates, the main substances that can adsorb MPs and NPs in the aquatic environment are suspended substances and sediments, with clay minerals, carbonate particles, as well as organic and inorganic matter in the substrate being the most important adsorption carriers.^24,53^ Studies have shown that MPs are the major contamination problem in the sediments of Bima City Bay, Indonesia,^54^ and similarly, the abundance of MPs in the sediments of the Ma River in Vietnam was relatively high (i.e., 1328.3 ± 1925.5 items/kg dry weight), while the abundance of MPs in the surface water was relatively low (i.e., 57.3 ± 55.8 items/m^3^).^55^ Under simulated seawater conditions, researchers observed strong heterogeneous aggregation of PSMPs with silicate clay minerals (kaolinite, bentonite, illitic and kaolinitic soil clays) and iron oxide minerals (hemeatite, goethite).^56^ Suspended substances and sediments etc. in the aquatic environment undergo heterogeneous aggregation with MPs and NPs through electrostatic adsorption and ligand exchange, which affects the distribution and diffusion of MPs and NPs in the aquatic environment,^57^ therefore it is necessary to investigate the differences in the adsorption effects of substances in the aquatic environment on MPs and NPs. Plastic particles of different particle sizes are affected by particulate matter in the aquatic environment and are transported and retained differently in the aquatic environment.^58^ Microplastics with larger particle sizes are more likely to be adsorbed and retained by natural sand in the aquatic environment,^59^ whereas nanoplastics with smaller particle sizes are less likely to be adsorbed by mineral particles.^57^ This may be due to the different surface curvatures of particles of different sizes, with smaller particles tending to have greater curvature than larger particles, resulting in poorer adsorption.^60^ However, if the size of the MPs is large, their adsorption may become less effective. When the particle size of polyamide microplastics (PAMP) is larger than 75 μm, it is difficult to remove by adsorption due to its large mass.^61^ Therefore, when plastic particles are adsorbed as adsorbents in the aquatic environment, the adsorption behavior that occurs in the aquatic environment varies greatly among plastic particles with different particle sizes.
There
are large amounts of natural colloids in the aquatic environment.
Natural colloids are colloidal particles with sizes between 1 nm and
1 μm, and are usually composed of natural substances, including
humus, proteins, polysaccharides and other natural organic matter
(NOM).^62^ The most studied natural colloids
are extracellular polymeric substances (EPS) and humic acids (HA).^63^ Natural colloids play an important role in the
behavior of plastic particles in the aquatic environment and in particular
play a key role in the stability of nanoplastics.^64^ Natural colloids have similar dimensions to NPs, making
it possible for NPs to heteroaggregate with natural colloids, whereas
the main behavior of MPs toward natural colloids is adsorption. MPs,
due to their large size, are mainly affected by their own gravity
for aggregation and sedimentation and are less affected by the process
of adsorption of natural colloids. After heteroaggregation of NPs
with natural colloids, an encapsulation layer of NPs covered by natural
colloids tends to form, which is an eco-colloid layer, also known
as an eco-corona.^65^ Eco-corona usually
shows colloidal stability, strong biosorption capacity and pollutant
transport capacity, which leads to altered aggregation and dispersion
of NPs and affects their stability in the aquatic environment.^66,67^ Since electrolytes in the aquatic environment interact with natural
colloids, the role of natural colloids on NPs is specifically analyzed
according to the level of electrolyte content (i.e., ionic strength)
in the aquatic environment.^68^ In the aquatic
environment that do not contain large amounts of electrolytes, the
presence of EPS and HA increases the stability of NPs in the aquatic
environment by increasing the steric hindrance and electrostatic repulsion,
making the NPs less likely to aggregate. It was shown that both HA
and EPS with PSNPs formed eco-corona significantly improved the transport
and stability of PSNPs in the aquatic environment, and the main mechanism
was that eco-corona increased the steric hindrance between the NPs
and the sand grains, which in turn facilitated the transport of the
NPs. The EPS, due to its larger molecular mass, would produce a stronger
steric hindrance, which also leads to its stronger transport enhancing
effect on PSNPs.^67,69^ However, when a large number
of electrolytes are present in the aquatic environment, for example
when a large number of electrolytes such as NaCl and CaCl2 are present in seawater, the effect of the natural colloid on the
NPs is somewhat altered due to the fact that the dissociation of the
electrolytes produces a large number of positive and negative ions.
The main mechanism is that under conditions of high ionic strength,
the electrical double layer of the natural colloid is compressed,
which is manifested as an increase in the ζ potential, leading
to easy aggregation of the NPs, while the bridging of divalent cations
with the natural colloid also makes the NPs easy to aggregate.^68^ The addition of Suwannee River humic acid (SRHA)
under simulated seawater showed that the critical coagulation concentration
(CCC) value of PET-NP was 98 mmol/L for Na^+^ at pH = 5.8,
and the CCC value of PET-NP decreased after the addition of different
molecular weights of SRHA in NaCl solution, implying that the stability
of NPs decreased and became vulnerable to aggregation. The smaller
molecular weight HA resulted in the greatest decrease in the CCC of
PET-NP, from 98 mmol/L to 53 mmol/L, while the larger molecular weight
HA introduced a higher steric hindrance along with a decrease in the
ζ potential of PET-NP, which partially counteracted the aggregation
effect.^70^
The formation of eco-corona also leads to changes in the ecotoxicity of NPs and their capacity to carry pollutants, which can alter the “Trojan horse effect” of NPs.^66^ The “Trojan horse effect” refers to NPs acting as carriers for other pollutants and transporting them.^71,72^ PSNP significantly enhanced the uptake and accumulation of phenanthrene in the gills and liver of rainbow trout, because smaller NPs have a greater potential for uptake and transport of hydrophobic organisms and show higher bioavailability in the presence of NOM, The uptake of phenanthrene was significantly reduced by about 70% because the presence of NOM competes with phenanthrene for adsorption on the surface of PSNP, reducing the transport of phenanthrene by PSNP and diminishing the “Trojan horse effect” of PSNP.^73^ The formation of eco-corona improved the biochemical response and environmental toxicity of NPs to varying degrees, depending on the composition of the NPs and the composition of the eco-corona.^74^ However, some researchers have shown that the phenolic hydroxyl group of HA can adsorb As(III) through hydrogen bonding, suggesting that the formation of eco-corona by NPs and HA may enhance the “Trojan horse effect” of As(III) adsorption.^75^ Therefore, we cannot directly conclude that the presence of NOM can reduce the ecotoxicity to aquatic organisms of NPs and the contaminants they carry, but we need to assess it according to the types of NOM, NPs and contaminants loaded.
The aging behavior
of plastic particles occurring in the aquatic environment is influenced
by a variety of factors, most notably light, mechanical stresses,
and biological factors. Biological aging involves the interaction
of plastic partlcles with microorganisms and will be discussed in
Section 6.2. Photoaging is one of the most important aging pathways
for plastic particles, leading to oxidation as well as changes in
surface crystallinity.^76,77^ Photoaging of plastic particles
is mainly achieved by stimulating the generation of reactive oxygen
species (ROS) on the surface of plastic particles by UV light, and
the possible mechanisms are shown in Figure 2 a.^78^ The chromophores
of plastic particles absorb photons to reach an excited singlet state
and form an excited triplet state by intersystem crossing. In the
aquatic environment, due to the presence of dissolved oxygen, the
excited triplet state can transfer energy to O2 to form
singlet oxygen (^1^O2). On the other hand, the
excited triplet state is able to transfer energy to the C–H
or C–C bond to generate alkyl radicals (R-), which can easily
react with O2 in water to generate peroxyhydroxyl radicals
(ROO−), which can capture hydrogen from the C–H bond
of the plastic particles to generate hydroperoxides (ROOH), and the
cleavage of the RO–OH bond leads to the formation of ·OH.^78,79^ Ultimately, these reactive oxygen radicals may be involved in the
photoaging process of plastic particles.^78,80^ Thus, the aging fate of light with plastic particles depends mainly
on the composition and structure of the plastic particles. Bare PSNPs
are exposed to light and produce ·OH and ^1^O2 and undergo aging behavior, but amino- or carboxyl-modified PSNPs
produce only a small amount of reactive oxygen radicals because of
the differences in the composition and structure of their surfaces.^76,78^ Due to their larger specific surface area, NPs are able to provide
more reaction sites and surface defects for UV absorption and radical
formation, whereas MPs are exposed to relatively few reactive oxygen
radicals due to their larger particles and their inherent high hydrophobicity
and chemical inertness. It was demonstrated that ·OH, O2^·–^ and ^1^O2 were found
in the aging of NPs, whereas only ·OH was detected in the aging
of MPs, and significant carbonyl functional group generation was detected
on aged PSNPs, whereas PSMPs showed no significant changes, and the
degree of aging of PSMPs was negligible compared to that of PSNPs.^78^ Scanning electron microscopy (SEM) studies revealed
that photoaging of PSNPs occurs uniformly on the surface, which is
attributed to the rapid generation of free radicals on the surface
of the PSNPs under UV irradiation and oxidation to produce small molecules,
that can be further mineralized to produce carbon dioxide. Changes
in surface morphology due to photoaging of MPs tend to be heterogeneous,
with portions receiving more light being more susceptible to surface
fragmentation and cracking, leading to spalling of submicroplastics.
Figure 2 Photo- and mechanical aging processes of MPs and NPs. (a) Processes and mechanisms of photoaging of MPs and NPs. Reproduced with permission from ref (78). Copyright 2022, Elsevier. (b). Fragmentation behavior of MPs during mechanical aging. Reproduced with permission from ref (84). Available under a CC-BY license. Copyright 2023, The Authors. (c). Aggregation behavior of NPs during mechanical aging. Reproduced with permission from ref (87). Available under a CC BY license. Copyright 2020, The Authors.
Photoaging is caused by the exposure of plastic particles to ultraviolet radiation, and other forms of radiation exist in the aquatic environment, such as thermal radiation.^81^ Thermal radiation aging of MPs in the aquatic environment is mainly due to the formation of new functional groups on the surface of MPs during the thermal process, which changes their physicochemical properties and leads to phenomena such as peeling of the outer skin of MPs.^82^ As there are different distribution levels of MPs and NPs in the aquatic environment, they are exposed to different degrees of thermal radiation. However, relatively few studies have been carried out on the aging of NPs by thermal radiation, and there is a need to focus on this aspect in the future.
Mechanical aging is likewise one of the most important ways of aging plastic particles, mainly in terms of changes in the size of the particles as well as changes in crystallinity, with water waves being the main source providing mechanical aging shear and tensile forces.^81,83^ In the case of MPs, the main environmental behavior during mechanical aging are the occurrence of molecular chain breaks and structural changes that result in fragmentation into smaller particles. The fragmentation behavior of MPs and NPs is shown in Figure 2 b,^84^ at an energy density level of 100 kJ/L, the 250 and 106 μm-sized pristine PSMPs were broken down into mean sizes of 120.6 ± 19.1 and 95.6 ± 16.8 nm, respectively. At an energy density of 4.0 kJ/L, shear stress resulted in the fragmentation of MPs with dimensions of 398 ± 54 nm into NPs with dimensions of 54 ± 25, 51 ± 15, and 73 ± 24 nm. At energy density levels higher than 40 kJ/L, the fragmentation of NPs was increased and NPs smaller than 10 nm in size were released.^85^ NPs, on the other hand, are often aggregated in addition to fragmentation to form smaller NPs, which is attributed to the increase in intermolecular attraction and electrostatic force between NPs and the decrease in surface energy during the mechanical aging process. And a layer of solvent or other impurities may be adsorbed on the surface of the NPs, resulting in a viscous effect. All these reasons lead to an increase in mutual displacement and collision between NPs and an increase in contact opportunities between particles, which promote the occurrence of aggregation.^85,86^Figure 2 c shows the aggregation behavior of NPs.^87^ It was found that the average hydrodynamic diameter of PSNPs with a diameter of 50 nm increased after 1 week of oscillation and the proportion of large PS particles increased.^88^ Dynamic light scattering (DLS) and SEM observed that 36.4 ± 12.6 nm NPs aggregated into 250 nm wide aggregation after mechanical oscillation,^85^ and these results suggest that slight aggregation of NPs occurs during mechanical aging.
Behavior
Leaching of contents is a common behavior in the aging process of plastic particles. It is important to be clear that there is no single component of any one type of plastic particle. Chemically, plastic particles are complex chemical mixtures of one or more polymers, supplemented by additives such as plasticizers, flame retardants, stabilizers, antioxidants, etc., containing unreacted monomers, unintentionally added impurities and byproducts.^89,90^ Since these substances are not covalently bound to the polymer, this means that plastic particles will gradually leach out their own additives when present in the aquatic environment.^91,92^ Common leaching substances include bisphenol A (BPA) and phthalates (PAE) used as plasticizers, which make plastic particles a persistent source of a wide range of organics in the aquatic environment. Plastic bags made of PE and PVC have been experimentally shown to leach PAE into the environment,^93^ and computational simulations have similarly illustrated leaching of additives from a range of polyolefin materials into the water, so it is not difficult to infer that MPs and NPs will similarly leach their contents into the environment.^94,95^ There are a number of factors that affect the leaching of additives from plastic particles, and the size of the plastics is one of the most important. This is because smaller plastics have a larger specific surface area and shorter diffusion paths, which means that the additives of smaller plastics can be leached through a greater number of choices and over shorter distances, making it easier to leach the additives.^96,97^ For leaching of di-n-butyl phthalate (DBP) from PVC plastics with different particle sizes, the equilibrium concentration and the pseudo-first-order model rate constant of leaching were both lower for large-size PVCMPs with a diameter of 5 mm compared to small-size MPs with a diameter of 100 μm, which suggests that the contact of plastic particles with water at a higher specific surface area will further lead to polymer dissolution and provide more sites for DBP leaching.^98^ NPs in seawater undergo swelling which causes fragmentation and polymer leaching, and poly(methyl methacrylate) (PMMANPs) exhibit more polymer leaching compared to PVCMPs.^99^ By assuming that the additives leaching process of polymers follows Fick’s law, which is consistent with Fick’s first and second laws, it can be inferred that the smaller the particle size, the faster the kinetics of additives leaching. Compounds with effective diffusion coefficients of 10^–16^ m^2^/s can be completely leached from plastic particles with radius of 100 nm or smaller in 1 h, while leaching in plastic particles with radius of 100 μm or larger is negligible.^100^ However, the existing experimental studies on the leaching of additives from NPs are not inadequate, and we cannot directly conclude that the same type of NPs have a stronger ability to leach additives than MPs. Based on the existing effects of MPs particle size on leaching ability and computational simulations, we hypothesize that the leaching ability of NPs additives is stronger than that of MPs. This means that NPs may provide higher doses of leached compounds locally.^101,102^
Plastic particles in the aquatic environment are often ingested by aquatic organisms that mistake them for food and subsequently undergo a range of post ingestion behavior. In general, NPs are more biotoxic than MPs, partly because NPs of small particle size penetrate cells and tissues more easily,^103,104^ and partly because the high specific surface area of NPs increases their bioavailability and contaminant carrying capacity.^105^ This subsection explores the differences in the behavior of MPs and NPs after ingestion by aquatic organisms from the level of individual organisms to the level of biological subcellular organelles. The behavior of MPs in fish causes only changes at the individual level, and does not lead to damage at the cellular or organelle level. It has been reported that the ingestion of MPs can cause a feeling of satiety in fish after ingestion, which is because after MPs are ingested by fish, due to the polymer structure of MPs, it is difficult for them to be decomposed by the digestive enzymes and digestive fluids of fish, resulting in MPs occupying the space of the fish’s stomach. MPs occupy space in the fish’s stomach and have a padding effect, causing a feeling of fullness and a decrease in the desire to eat.^106,107^Figure 3 a demonstrates the occupation of fish stomach space by microplastics.^108^ The results of a number of researches have shown that MPs do not accumulate in the digestive tract of fish for long time, but rather travel through the entire digestive tract and are excreted in faecs.^109,110^ Ingested MPs also cause changes in the structure and composition of the fish gut microbial community. For example, levels of mucins and D-lactate in the gut of marine medaka ingesting PSMPs were significantly increased, suggesting an increased proportion of bacteria and impaired immune function in the gut of marine medaka.^108^ Similarly, nanoplastics can lead to changes in the gut microbiota of exposure to environmentally significant and sublethal levels of 80 nm PSNPs was found to alter the composition of gut microbial communities in grass carp, affecting diversity and resulting in gut microbial imbalances,^111^ whereas reports of nanoplastics causing satiation in fish are rare. On the one hand, it is because NPs are very small in size and hardly occupy the stomach space of fish, and on the other hand, the more important reason is that NPs usually do not stay in the stomach of fish but break through the intestinal barrier of fish to enter into other organs of the fish’s body to cause deeper damage, which is shown in Figure 3 b.^112^ Observation of the particle size distribution of polyamide 66 (PA66) in rat feces after ingestion of PA66 revealed that PA66-MPs could be completely excreted, whereas approximately 30% of the ingested PA66-NP were not detected, suggesting that PA66-NPs could cross the intestinal barrier and enter the blood circulation.^113^ PSNPs ingested by goldfish effectively crossed the intestinal barrier and did not remain in the digestive tract, but reached the internal liver and muscle tissues and induced histopathological changes after distribution through the bloodstream.^114,115^
Figure 3 Differences in the behavior of MPs and NPs after ingestion by aquatic organisms. (a). The occupation of fish stomach space by microplastics. Reproduced with permission from ref (108). Copyright 2020, Elsevier. (b). The breaking through the fish intestinal barrier of NPs. Reproduced with permission from ref (112). Available under a CC-BY license. Copyright 2020, The Authors. (c). Oxidative stress induced by nanoplastics via ROS. Reproduced with permission from ref (127). Copyright 2023, Elsevier. (d). Aggregation of microplastics after excretion. Reproduced with permission from ref (129). Copyright 2019, Elsevier.
NPs can enter the cell interior and cause damage at the organelle and suborganelle levels, leading to cytotoxicity and even genotoxicity.^116^ The entry of NPs into cells stems from their adsorption to the outer cell membrane, and these NPs can activate the cells to enter the cell interior by endocytosis or passive diffusion.^112,117^ Then, these NPs internalized into the cells further reach various organelles. PSNPs can accumulate in the mitochondrial lipid and disrupt mitochondrial function and integrity by interfering with the production of adenosine triphosphate (ATP) and nicotinamide adenosine dinucleotide phosphate (NADP) and by inducing the production of peroxides in mitochondria.^118,119^ The process of intracellular transfer of NPs is complex, but their final destination is lysosomal deposition by intracellular transport.^120,121^ Different types of PSNPs can be transferred to lysosomes via clathrin-mediated or caveolae-mediated pathways.^122^ More importantly, NPs can reach the nucleus causing DNA damage and genotoxicity. Short-term exposure to a concentration of 100 nm PSNPs induces DNA damage in Mytilus galloprovincialis and affects the expression of genes related to immune functions.^123^ Prolonged exposure to the 44 nm PSNPs causes DNA damage in goldfish erythrocytes, which is manifested by an increase in erythrocyte nuclear abnormalities (ENAs) and induces permanent damage.^114^
In addition to the differences in distribution in the fish, the differences between MPs and NPs in causing oxidative stress were also significant. Despite the hypothesis that MPs cause oxidative stress in fish, no strong evidence has actually been found to support that ingestion of larger PSMPs induces oxidative stress mediated by antioxidant defense-related genes.^124^ Animal experiments have similarly shown that MPs induce no or minimal adverse effects related to oxidative stress.^125,126^ On the contrary, induction of oxidative stress, a potential toxic mechanism of NPs, is caused by excessive ROS, which lead to oxidative damage in organisms and alter activities of enzymes in organisms, which is shown in Figure 3 c.^127^ A decrease in the activities of major antioxidant enzymes were observed in medaka after ingestion of NPs, indicating that these enzymes gradually lost their ability to resist ROS and thus the enzyme activity was inhibited after prolonged exposure to NPs, and these phenomena suggest that NPs can induce oxidative stress in medaka and affect its antioxidant system.^127^
The process of excretion of MPs and NPs from fish is also very different, with MPs being excreted with feces after a period of time in the fish’s stomach because they do not cross the fish’s intestinal barrie.^128^ As the MPs are coated by the intestinal mucus during the excretion process, this liquid is denser than water, so the excreted MPs have a stronger sinking and aggregation behavior in water compared to the original particles, which alters the vertical movement behavior of MPs in the aquatic environment, which is shown in Figure 3 d.^129^ While studies on whether NPs can be excreted from the body have not been reported in detail, it is clear that the retention time of NPs in the fish body is significantly longer than that of MPs, and that NPs accumulate in the liver and muscle tissues of the fish, and even transport throughout the body with the blood circulation.^130^
The interaction of plastic particles with microorganisms is to some extent the biological aging process of plastic particles. Biological aging is the aging process that occurs when plastic particles are exposed to biologically active substances in the environment and the proliferation of organisms on the surface of plastic particles. The main biologically active substances in the aquatic environment are microorganisms and some plants and animals and biological enzymes. These bioactive substances can change the physical and chemical properties of plastic particles, inducing surface oxidation, reducing their hydrophobicity, etc.^131^ The surface of pristine MPs is relatively smooth, with relatively little direct contact and interaction with organisms, so the colonisation process of microorganisms on MPs is relatively time-consuming, which leads to a very slow process of biological aging of MPs. Over time, a large number of microorganisms colonize the MPs eventually forming biofilms. Algae adhered to the surface of PEMPs during the 30 days biological aging process and biofilms were formed on the surface of the MPs. The formation of biofilm changed the surface morphology and wettability of the plastic.^131^ MPs can be provided to these microorganisms as a carbon source, leading to accelerated aging of MPs, which is particularly evident in degradable plastics. Comparison of biodegradable and nonbiodegradable poly(butylene succinate) (PBS) and polylactic acid (PLA) MPs biological aging processes revealed that degradable PBS and PLA MPs particles produced relatively more secondary MPs than nondegradable MPs particles.^132^ Comparison of the biological aging process of degradable and nondegradable MPs in the ocean showed that biodegradable MPs have a higher bacterial colonisation rate compared to nondegradable MP.^133^ NPs, on the other hand, have a large specific surface area and are full of defects, making them more susceptible to interaction with microorganisms and increasing the biological aging of NPs. At a high ionic strength, 20 nm NPs could attach to the cell surface, while no significant direct attachment of 2 μm MPs with bacteria was observed.^134^ However, since most of the NPs are smaller than microorganisms, this causes the microorganisms to be difficult to colonize the NPs; instead, the NPs attach to the microorganisms, which prevents the formation of biofilms during the biological aging process of the NPs. When multiple NPs are attached to a single microorganism, EPS has an adhesive effect that alters the aggregation morphology of NPs.^135^ PP, PE, and PVCNPs were able to attach onto Lactobacillus cells through electrostatic interactions, van der Waals forces, and hydrogen bonding, while extracellular proteins on the surface of Lactobacillus were also observed to polymerize with NPs.^136^Figure 4 demonstrates the adsorption of PSNPs on microorganisms and the formation of biofilms on PSMPs.^137^
Figure 4 Adsorption of PSNPs on microorganisms and the formation of biofilms on PSMPs. (a). Adsorption of PSNP on E. coli. (b). Adsorption of PSNP on Bacillus sp. (c). Formation of PSMP-biofilm complex. (d). EPS secreted during PSMP biofilm complex formation. Reproduced with permission from ref (137). Copyright 2022, Elsevier.
There is growing evidence that biofilms affect the environmental behavior of MPs by altering their physicochemical properties, which is shown in Figure 5 a.^138^ For example, the increasing thickness of the biofilm leads to an increase in the size and density of the MPs, thus altering their ability to move vertically in the water.^139,140^ The ability of MPs forming biofilms to accumulate substances in the aquatic environment also changed. This is because the EPS secreted by microorganisms after the formation of biofilm leads to a relative increase in the number of accessible sites on the surface of MPs, and the formation of biofilm increases the number of oxygen-containing functional groups on the surface of MPs, which enhances the bonding energies of the hydrogen and Π–Π bonds, thus leading to the accumulation of a greater amount of environmental substances.^141^ For example, biofilm-forming MPs enhanced the accumulation of metal ions in the aquatic environment compared to pristine MPs without biofilm formation. Biofilms not only enhanced the adsorption capacity of MPs for metals, but also enhanced the stabilization of metals on MPs.^142,143^ Microbial communities that form biofilms on MPs are more specialized compared to those that aggregate on natural substrates. This is because the process of microbial biofilm formation on the surface of MPs is actually a process of sorting microbial species, and the microbes that are ultimately able to colonise the MPs are those that are able to metabolize the MPs as a substrate.^144^ Higher abundance of microorganisms was observed on MPs substrates compared to natural substrates. This is mainly due to the fact that MPs provide colonisation sites for microorganisms while also adsorbing organic matter from the aquatic environment, which can be a source of nutrients for microorganisms.^145^ Although higher microbial abundance was observed on MPs substrates, this does not mean that MPs are beneficial to microorganisms. The microbial community structure and physiological functions in MP biofilms differ from those of natural substrate biofilms,^146,147^ and biofilms formed with MPs as substrates have equally important effects on the colonised microorganisms. Due to the presence of MPs, the biodiversity of biofilm systems may be altered and the proportion of certain species may be suppressed by MPs,^148^ MPs may also lead to potential changes in the carbon and nitrogen metabolism of biofilms,^149−151^ and the unique bioecological niche of MP biofilms has even been suggested to be a reservoir for certain microorganisms such as pathogens, antibiotic resistant bacteria (ARB) and antibiotic resistant genes (ARGs).^152^
Figure 5 Effect of biofilm formation on MPs and the effect of NPs on microorganisms. (a). The effect of NPs on microorganisms. Reprinted with permission from ref (138). Copyright 2021, ACS. (b). Effects of nanoplastics on microbial gene expression. Reprinted with permission from ref (157). Copyright 2023, Royal Society of Chemistry.
The most common effect of NPs on microorganisms is that they can inhibit microbial growth processes and have a significant inhibitory effect on the chlorophyll synthesis process of microorganisms such as algae.^153^ NPs may also cause deformation and damage of the cell wall of bacteria and can cross the cell membrane into the bacterial interior by means of membrane permeation.^154,155^ Due to the ability of NPs to enter the interior of microbial cells, this leads to NPs causing DNA damage and affecting processes such as photosynthesis and proliferation of algae.^156^ In addition, NPs can affect microbial gene expression and promote the splicing transfer of antibiotic resistance genes binding, shown in Figure 5 b, which may lead to a decrease in the effectiveness of antibiotics in the future.^157^
and Perspectives
This paper systematically elucidates the differences in the behavior of NPs and MPs in the aquatic environment and provides new insights to clarify the multifaceted differences between the two in the aquatic environment. The main differences in the behavior of MPs and NPs in the aquatic environment are summarized in Table 2. Considering the differences in the behavior of NPs and MPs in the aquatic environment are multifaceted, it is more important for researchers to distinguish between NPs and MPs. The following points should be focused on in future research.
This work was supported by the National Natural Science Foundation of China (Nos. 42073082, 12005284), the National Key Research and Development Program of China (2019YFC1805300).
The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/envhealth.4c00013.
^‡^ S.W. and N.S.A.-H. contributed equally to this work.
The authors declare no competing financial interest.