Micro- and nanoplastics’ transfer in freezing saltwater: Implications for their fate in polar waters

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Focusing on nanoplastics

From the previous hypotheses, it can be assumed that fragmentation of plastics to micro-metric and submicrometric sizes is one of the several fates of plastic debris. Based on the diversity of environmental mechanisms that can incidentally produce small microplastics (1 to 1000 µm) and nanoplastics (<1 µm) (e.g.: mechanical abrasion, photo-oxidation, etc.), these plastic particles are expected to be present in all environmental compart-ments. Therefore, it is essential to assess their environmental fate in order to i) resolve the mass balance of plastic debris, ii) determine organisms’ level of exposure and iii) de-termine their potential impact on Earth system processes. However, due to their smaller size, nanoplastics have colloidal properties that make them significantly diﬀerent from larger particles in several respects and warrants studying them separately.
A colloidal dispersion is a system where one phase (liquid, solid or gas) is dispersed in a diﬀerent continuous phase. In our case, solid nanoplastics particles are dispersed in liquid (Goodwin 2004; Hiemenz and Rajagopalan 1997). There have been some attempts to give a clear-cut definition of colloidal dispersions, such as the CRC Handbook of Chemistry and Physics definition: molecules or polymolecular particles dispersed in a medium that have at least in one direction a dimension roughly between 1 nm and 1 µm(Haynes, Lide, and Bruno 2015). However, it will become clear that colloidal properties which allow particles to remain dispersed within another medium are highly dependent on the physicochemical properties of the continuous phase.
Due to their colloidal properties, nanoplastics may have a diﬀerent environmental fate compared to larger particles. Furthermore, studying them requires the use of various analytical methods and theoretical frameworks compared to larger particles. Indeed, as particles’ size decreases down to the colloidal size range, they transition away from motion dictated by gravitational forces and towards motion dictated by intermolecular forces (Elimelech 1998; Goodwin 2004; Hiemenz and Rajagopalan 1997) . The particle size at which this transition occurs depends on the relative densities of the dispersed phase (particle) and of the continuous phase (liquid) and the liquid viscosity, as defined by Stokes’ law (Stokes 1851), as well as the particle size as defined by the Stokes-Einstein equation (Einstein 1905; Sutherland 1905). Indeed, Stokes’ law :
shows that the settling speed of a spherical particle Vs (m s−1) is a function of the diﬀerence between its density ρp and the density of the fluid ρf (kg m−3), as well as the dynamic viscosity of the fluid µ (kg m−1 s−1), the square of the particle radius rp (m) and gravitational acceleration g (equal to 9.8 m s−2 on Earth). Furthermore, the collision of small particles with water and solutes causes their hydrodynamic diﬀusion, also called Brownian motion. This is illustrated in the Stokes-Einstein equation:
which relates a spherical particles’ diﬀusion coeﬃcient D (m2 s−1) to its radius, the fluid’s viscosity and the thermal energy of agitation, given by the Boltzmann constant kB (kg m2 K−1 s−2) and the temperature T (K).
While Equation 1.1 shows that the settling speed Vs is proportional to the size (r2p), Equation 1.2 means that the Brownian motion, characterized by D, is inversely propor-tional to its size. Therefore, when the size decreases D becomes predominant compared to Vs. Based on these equations, carbon-based particles in aqueous systems are deemed to be colloidal around 1 µm, which is why nanoplastics are defined as submicrometric. This Brownian motion is an important consideration when assessing environmental transport as well as during their analysis and theoretical study. For example, due to Brownian mo-tion, nanoplastics cannot be extracted from environmental media with the density-based methods used for microplastics.
Colloids are also characterized by a high specific surface area, defined as the total sur-face area per particle mass. This renders surface interactions crucial in shaping colloidal behavior and in selecting appropriate methods of analysis. These surface interactions (e.g.: electrostatic repulsion and Lifshitz van der Waals attraction) operate at short dis-tances from the particle’s surface (up to approximately 50 nm)(Israelachvili 2015). The section on the Derjaguin-Landau-Verwey-Overbeek (DLVO) theory of colloidal stability (cf : Section 1.2.2) will provide a review of the surface interactions. Given a favorable (at-tractive) balance of surface interactions and hydrodynamic forces, colloids sorb onto other species (e.g.: other colloids, molecules, surfaces). The properties of species onto which nanoplastics sorb (or that sorb onto nanoplastics) significantly modifies nanoplastics’ overall physicochemical properties, such as their dimensions, surface chemistry, etc. Fur-thermore, nanoplastics’ size may be comparable to that of environmental macromolecules. Therefore their sorption onto these molecules may strongly modify nanoplastics’ physic-ochemical properties.
A final consideration that sets nanoplastics are apart from microplastics is that diﬀer-ent optical methods must be used when analyzing microplastics and nanoplastics. Indeed, since nanoplastics’ size is similar to the wavelengths of visible light, they cannot be de-tected by optical instruments which are diﬀraction-limited (e.g.: light microscopy and infrared spectroscopy) (Gigault, Halle, et al. 2018; Gigault, El Hadri, Nguyen, et al. 2021).
Nanoplastics are a contaminant of emerging concern (CEC) since they are « new com-pounds or molecules that were not previously known or that just recently appeared in the scientific literature » (Sauvé and Desrosiers 2014). The impacts of this emerging contam-inant on organisms and environmental processes have become a global concern for the public and policymakers (Allan, Sokull-Kluettgen, and Patri 2020; GESAMP 2015; SA-PEA, Science Advice for Policy by European Academies 2019). Therefore, nanoplastics have been the topic of an increasing amount of scientific investigation (Alimi, Farner Bu-darz, et al. 2018; da Costa et al. 2016; Lehner et al. 2019 and references therein). The focus of this work is to determine nanoplastics’ environmental fate to assess their poten-tial risk. However, it is still unclear whether nanoplastics may be a hazard since studies investigating their eﬀects on ecosystem and human health have rarely used nanoplas-tic particles that are representative of nanoplastics found in the environment. Instead, studies have often used model nanoplastic particles composed of PS and suspended with additives such as preservatives, antimicrobials, or surfactants. Pikuda et al. demon-strated that the (eco)toxicity of nanoplastics was usually caused by the additives added to the liquid dispersion of nanoplastic models rather than the plastic itself (Pikuda et al. 2019). Conversely, nanoplastics are expected to be (eco)toxic in large part due to the leaching of additives added during the manufacturing process (e.g.: brominated flame retardants, phthalate plasticizers, and lead heat stabilizers) and the release of (eco)toxic monomers (e.g.: PUR, and polyacrylonitrile) (Lithner, Larsson, and Dave 2011). How-ever, to date (eco)toxicity studies have mainly focused on pristine polystyrene particles that are free of additives used during manufacturing. Nanoplastics may cause deleterious eﬀects other than (eco)toxicity, for example, by impacting ecosystem processes, such as biogeochemical cycling (L. Galgani and S. A. Loiselle 2020). Therefore, a One Health perspective, combining transdisciplinary studies in the domains of human, animal and environmental health is called for (Prata et al. 2021).

Approaches to assess the environmental fate of nanoplastics

The role of experimental approaches

Assessing the transport and accumulation of nanoplastics in the environment benefits from previous approaches developed for engineered nanomaterials (ENM), since these are also colloidal anthropogenic contaminants (Dale et al. 2015; Mitrano, Wick, and Nowack 2021; Gigault, El Hadri, Nguyen, et al. 2021). A preliminary step consists in conceptualizing the life cycle of nanoplastics. Sources of nanoplastics are multiple since they can be generated from plastic objects and plastic debris by various mechanisms.
So, to assess the transport and accumulation pathways of nanoplastics in the environment, it is crucial to identify the natural physical and chemical processes that may impact their behavior in all environmental compartments (e.g.: freshwater, seawater, soils, etc.).
While, rejection of plastic aerosols to the atmosphere can occur (Allen et al. 2019; Bergmann, Mützel, et al. 2019; Wik and Dave 2009), current evidence shows that most plastic debris is water-bound (Horton et al. 2017; Schwarz et al. 2019 and references therein). Therefore, the main transport pathways of nanoplastics in the environment are through aqueous systems. Based on the current understanding of the transport and accu-mulation pathways of ENM’s, the most relevant abiotic processes aﬀecting nanoplastics’ fate in the environment are summarized in Figure 1.3. Two key processes are nanoplas-tics’ ability to aggregate with other particles, and their ability to be stabilized by the sorption of dissolved (i.e.: low molecular weight) organic species ((Wang et al. 2015; Yu, Jingfu Liu, et al. 2018) and references therein). Indeed, nanoplastics are more likely to hetero-associate with naturally occurring species than to homo-aggregate with other nanoplastic particles, since the concentration of natural dissolved or particulate species in aqueous environmental systems are in the range of µg L−1 to mg L−1 (Benner 2002; Burdige 2002; Sanderman, Baldock, and Amundson 2008; Stumm and Morgan 1996), whereas projected nanoplastic concentrations are in the range of pg L−1 to µg L−1 (Lenz, Enders, and Nielsen 2016). Homo-aggregation of a nanoplastic with another nanoplas-tic could occur at the surface of a disintegrating piece of plastic debris, where surface concentrations of nanoplastics may be high.
The impact of hetero-association on nanoplastics’ mobility is highly dependent on the species’ molecular weight. For simplicity, natural species are classified as either dissolved (e.g.: electrolytes, dissolved organic matter, etc.) or particulate (e.g.: metal oxides, clays, particulate organic matter, etc.) Even though the transition between the two is inherently blurry, it has often been fixed at arbitrary cut-oﬀ sizes (mostly 0.2 and 0.45 µm) due to the widespread use of filters as separation methods. Truly dissolved species are under a few kDa in molecular weight and in thermodynamic equilibrium with the water. Here, for the purpose of this work which aims at understanding plastic colloids’ environmental fate, large macromolecules that pass through 200 nm filters and have molecular weights up to approximately 106 g mol−1 are considered dissolved. The dimensions of particulate species can be as small as a few nanometers but they are orders of magnitude denser than dissolved species.
Hetero-aggregation of nanoplastics with particulate species can produce aggregates that either remain dispersed, cream (move to the surface) or settle, depending on their size, shape, density, and porosity. Stokes’ law (Equation 1.1) presents the eﬀects of particle size and density, assuming particles are spheres. However, aggregates cannot be simply considered as larger spherical particles since they have a nonspherical shape and high porosity, which impacts their buoyancy. For example, colloidal aggregation can give rise to aggregates with a linear structure (i.e.: low fractal dimension ≈1.6) in conditions where colloids immediately adhere to each other when they first come into contact. However, in conditions that are less favorable to adhesion, aggregates have a more compact structure (i.e.: high fractal dimension ≈2) (Hackley and M. A. Anderson 1989). The more linear (less compact) aggregates tend to settle more rapidly due to the reduced drag force of the fluid on the aggregate, compared to a permeable sphere of equivalent density (C. P. Johnson, X. Li, and Logan 1996). Aggregation not only reduces nanoplastics’ colloidal properties but also aﬀects other (downstream) processes, such as deposition in porous media (e.g.: soils, aquifers, sediments, etc.)(S. Lin and Wiesner 2012a) and bioavailability (Lebordais et al. 2021), etc.
Another critical process is nanoplastics’ ability to sorb dissolved matter onto their surface. This layer of sorbed materials, called eco-corona, provides nanoplastics a new type of identity (i.e.: surface charge composition, optical property, etc.) (Wheeler et al. 2021). The eco-corona can be composed of naturally occurring molecules in solution, such as dissolved organic matter, exudates from microorganisms, or even dissolved contami-nants. The formation of this eco-corona and its ability to stabilize nanoplastics against aggregation depends on the water’s electrolytic composition, as well as the chemical com-position of the molecules and of the nanoplastic (Buﬄe et al. 1998; Yu, Jingfu Liu, et al. 2018).
Once the abiotic processes that control nanoplastics’ fate have been identified, assess-ing nanoplastics’ environmental fate requires the iterative use of three complementary approaches, as depicted in Figure 1.4:
– One approach consists in undertaking experiments that model nanoplas-tics’ transport pathways in environmental systems to identify and quan-tify the processes that control nanoplastics’ environmental fate. Two complementary types of experimental systems exist. First, there are those that attempt to elucidate possible outcomes by measuring probabilities that specific processes occur or the rates at which these occur. In this case, experiments rapidly and empirically define transport rates and can give insights into probable mecha-nisms. For example, colloids are observed to be retained in soils. This is potentially due to diﬀerent processes (e.g.: physical entrapment in soils or chemical aﬃnity for surfaces), but these processes are not identified (Hendren et al. 2015). The second type of experimental systems attempt to elucidate the mechanisms behind a given behavior. This consists in simplifying the system until a given parameter is iso-lated and a mechanism identified. For example, this could ascertain that colloids are retained due to chemical aﬃnity for soil surfaces.
– Another approach consists in numerically modeling nanoplastics fate and behavior by combining the rates and probabilities obtained experimentally with data on nanoplastics’ presence and/or concentration. For nanoplastics, this data can be estimated by MFA, plastic degradation rates, and extrapolated from mi-croplastic concentrations (Albert A Koelmans et al. 2017a; Lenz, Enders, and Nielsen 2016). Numerical simulations can either predict the environmental fate of nanoplastics (Besseling et al. 2017) or model all possible outcomes to determine the key processes controlling nanoplastics’ fate (Clavier, Praetorius, and Stoll 2019).
– Finally, numerical simulations and experimental results can be confronted with data on environmental presence or concentrations of nanoplastics obtained from the characterization of field samples to assess their valid-ity. Inversely, results from the characterization of field samples can give informa-tions on the processes that must be investigated by experimentation and numerical simulations. For example, if high concentrations of nanoplastics are quantified in soils, numerical and experimental models should focus on quantifying deposition rates as a function of soil properties to quantify exposure to soil-dwelling organisms and potential impacts on soil properties.
This tiered approach is an iterative approach that has been used successfully to study nat-ural, engineered, and incidental particles. However, diﬀerent challenges arise depending on the nature of the particle.
When studying natural colloids, information about colloidal fate can be obtained from appropriate sampling and characterization of field samples, complemented by experimen-tal and numerical modeling studies (Buﬄe et al. 1998; Filella 2007 and references therein). Field sampling is not always the most suited approach for engineered colloids, such as ENM, since their release in the environment occurs as occasional point-source contami-nation events or diﬀuse contamination in low concentrations(Bystrzejewska-Piotrowska, Golimowski, and Urban 2009). Therefore, the goal has been to design environmentally relevant laboratory experiments to study the diﬀerent ENM transport and transforma-tion pathways they can undergo in the environment (Peijnenburg et al. 2015). Other challenges occur for incidentally produced colloids, such as nanoplastics, depending on the nature of the colloid and their location. If present in high concentrations (e.g.: from point sources) and/or easily distinguishable from the biogeochemical background (e.g.: mineral oxides in mines and soot), characterization of field samples can be a suited ap-proach. However, this approach is technically unfeasible for nanoplastics since they are both carbon-based particles and present in low concentrations compared to background concentrations of carbon contained in natural organic matter (NOM). Therefore, since nanoplastics are challenging to detect in natural samples, studying their environmental fate in laboratory experiments is the most appropriate approach.
Contrary to ENM that are manufactured to have specific properties, nanoplastic parti-cles are expected to have heterogeneous properties (i.e.: a variety of shapes, compositions, surface properties, etc.)(Gigault, El Hadri, Nguyen, et al. 2021). Therefore, to prop-erly assess nanoplastics environmental fate, another challenge, besides choos-ing environmentally relevant experimental conditions, is to study environ-mentally relevant nanoplastic models. In conclusion, evaluating nanoplastics’ fate in the environment today must rely on numerical models of nanoplas-tic transport underpinned by experimental research. In parallel, analytical techniques to characterize them in environmental samples must be developed.

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Global theoretical frameworks

Theoretical frameworks are necessary to interpret experimental results and to implement numerical models. In the following section, two theoretical frameworks commonly used to assess the fate of colloids will be described: 1) particle collision rates and attachment eﬃciencies, and 2) the Derjaguin-Landau-Verwey-Overbeek (DLVO) theory of colloidal stability and its extended versions.
Particle col lision rate (β) and attachment eﬃciency (α)
To elucidate the environmental fate of colloids such as nanoplastics, it is necessary to break down their behavior into two independent steps: 1) the rate of colloids’ collision with a surface (of another particle or a larger object) and 2) the probability that this collision results in attachment. These two steps can be treated independently since the forces of colloidal attraction that lead to attachment are of a shorter range than the particle size. Therefore they do not aﬀect collision rate. Figure 1.5 illustrates the major processes and parameters of the solution that impact particle collision rate and attach-ment eﬃciency. They are summarized in the empirical predictors β: particles’ collision rate with other particles or surfaces and α: particles’ attachment eﬃciency (Elimelech 1998; Xing, Vecitis, and Senesi 2016).

Table of contents :

1 Assessing the environmental fate of nanoplastics: A critical review of aggregation processes
1.1 Nanoplastics: what and where?
1.1.1 Plastic debris as environmental contaminants
1.1.2 How a revolutionary material became an environmental concern
1.1.3 Transport and transfer of plastic debris
1.1.4 Focusing on nanoplastics
1.2 Approaches to assess the environmental fate of nanoplastics
1.2.1 The role of experimental approaches
1.2.2 Global theoretical frameworks
1.3 Nanoplastic stability in water
1.3.1 Nanoplastic models
1.3.2 Solution composition
1.3.3 Sample preparation methods
1.3.4 Instruments and methods to assess the stability
1.3.5 Interpretation in light of theoretical frameworks
1.4 Nanoplastics transport and retention in interfaces of the hydrosphere
1.4.1 Solid/Liquid interfaces of continental systems: porous media
1.4.2 Solid/Liquid interfaces of polar systems: sea ice
2 Stabilization of fragmental polystyrene nanoplastic by natural organic matter: Insight into mechanisms
2.1 Introduction
2.2 Experimental section
2.2.1 Sample Preparation
2.2.2 Size characterization
2.2.3 Kinetics of Colloidal Aggregation
2.2.4 Derjaguin Landau Verwey Overbeek (XDLVO) theory of colloidal stability
2.3 Results and Discussion
2.3.1 Colloidal stability of nanoplastic models
2.3.2 Stabilization of NPT-P by natural organic matters
2.3.3 Colloidal stability of NPT-P according to the nature and concentrations of NOM
2.3.4 Environmental Implications of NOM-NP interactions
2.4 Conclusion
3 Deposition of environmentally relevant nanoplastic models in sand during transport experiments
3.1 Introduction
3.2 Methods
3.2.1 Dispersions of nanoplastic models
3.2.2 Charge characterization
3.2.3 Size characterization
3.2.4 Transport in porous media
3.2.5 Theory
3.3 Results and Discussion
3.4 Conclusion
4 Deposition of nanoplastics: The roles of size polydispersity and natural organic matter
4.1 Introduction
4.2 Materials and Methods
4.2.1 Materials
4.2.2 Methods
4.3 Results and Discussion
4.4 Conclusion
5 Micro- and nanoplastics’ transfer in freezing saltwater: Implications for their fate in polar waters
5.1 Introduction
5.2 Materials and Methods
5.2.1 Materials
5.2.2 Methods
5.3 Results and Discussion
5.4 Conclusion
5.5 Supplementary Data
6 Conclusion and Perspectives