Nanoplastics: what and where?
Plastic debris as environmental contaminants
Plastic debris has contaminated all of Earths’ compartments, from freshwater lakes, rivers, and sediments (Klein, Worch, and Knepper 2015; Eriksen et al. 2013), soils (Scheurer and Bigalke 2018; Zubris and Richards 2005), seas and oceans that are close to anthropogenic activity (Thompson 2004), and more remote (Lusher et al. 2015), as well as some of the deepest ocean sediments (Peng et al. 2020), the atmosphere (Allen et al. 2019) and biosphere (L. Li et al. 2020; Provencher et al. 2019 and references therein). Due to this planetary-wide contamination, plastics’ environmental fate can now be studied through the lens of a plastic cycle, comparable to the carbon or nitrogen cycle (Bank and S. V. Hansson 2019). Furthermore, plastics are now proposed as a marker of the Anthropocene, a proposed geological era defined by widespread anthropic activity (S. L. Lewis and Maslin 2015; Waters et al. 2016).
Contamination is defined here as the introduction of a foreign object or energy into the environment, whereas pollution is contamination that causes harmful eﬀects to ecosys-tems and/or human health (GESAMP 1991). As for other environmental contaminants, to define the magnitude of the risk posed by plastics, two sides of an equation must be solved (Albert A. Koelmans et al. 2017b):
1. their concentration in an area, which represents organisms’ level of exposure and their probability of coming into contact with environmental processes
2. their hazard to a given organism or environmental process.
Therefore, to better understand how plastic debris may cause harm, it is essential to study how they are transformed, transported, and accumulated in order to shed some light on their environmental concentrations. However, understanding plastics’ environmental fate is complicated due to i) the numerous sources of plastic debris into the environment, ii) the long-range dispersal of plastic debris by air and water currents caused by plastics’ lightweight and relative durability and, iii) the incidental production of plastic particles from larger plastic objects.
How a revolutionary material became an environmental concern
Plastics are synthetic water-insoluble materials that are produced by chemically linking the same simple organic molecules called monomers, into polymers with high molecular weight. The development of organic chemistry and industrial processes in the 1940s and 1950s has allowed the development and mass commercialization of plastics with a wide array of modular properties (Andrady and Neal 2009). Plastics generally belong to one of these functional categories:
– thermoplastic polymers, which can be hardened and softened repeatedly by tem-perature changes;
– thermosetting polymers, which when cured by heat or other means become insoluble and unable to melt;
– elastomers, which can be deformed by mechanical stress and rapidly recover;
– polyurethane (PUR) resins, which can be composed of diﬀerent monomers linked with urethane. (International Organization for Standardization, 2013).
The following thermoplastic and thermoset polymers: polyethylene (PE), polypropylene (PP), polystyrene (PS), polyvinyl chloride (PVC), polyethylene terephthalate (PET) and fibers of polyester, polyamide, and acrylic (PP&A) constitute 89% of all the plastic poly-mers produced (Figure 1.1, inner circle). Finally, PUR resins, which are neither thermo-plastics nor thermosets, are also produced in high quantities and comprise approximately 7% of the total polymer production (Figure 1.1, inner circle) (Geyer, Jenna R. Jambeck, and Kara Lavender Law 2017). Plastics’ properties depend on the monomers and addi-tives that compose the plastic, as well as the production process. Incorporating diﬀerent additives can confer a wide array of functionalities, such as improved plasticity during manufacturing and, increased rigidity, lightness, and resistance to fire(Hahladakis et al. 2018). Therefore, additives are a non-negligible component of plastics, with variable con-centrations according to the plastic composition and use (Geyer, Jenna R. Jambeck, and Kara Lavender Law 2017; Lithner, Larsson, and Dave 2011).
Figure 1.1: Mass-based distribution of worldwide plastics production (inner circle) and plastic waste generation (outer circle) in 2015, according to polymer type and additive. The inner circle corresponds to a total of 407 million metric tons (Mt) produced, and the outer circle corresponds to 302 Mt discarded. The primary market sector for each polymer is indicated, with C&I = Consumer and Institutional. (Source = Geyer, Jenna R. Jambeck, and Kara Lavender Law 2017). Figure produced with https://rawgraphs.io
Plastics have become an inherent component of the modern industrial world. They are generally produced from fossil carbon sources at an aﬀordable monetary price. These lightweight and durable materials have brought about significant societal advancements, for example, by increasing access to medical care and technology, by replacing heavier and more fragile alternatives, as well as slowly renewable or rare natural materials(Andrady and Neal 2009). Therefore, by 2015 it was estimated that 8.3 billion metric tons of plastic had been produced, which makes plastics the third most-produced material behind steel and cement (Geyer, Jenna R. Jambeck, and Kara Lavender Law 2017). To place plastics in a planetary context, the mass of plastic produced is estimated to be roughly twice that of all animals on Earth (Elhacham et al. 2020).
The increasing mass of plastic debris generated has been the unforeseen consequence of introducing such vast amounts of plastics in our livelihoods. Plastic debris is defined here as materials mainly composed of synthetic or semi-synthetic water-insoluble poly-mers that are present in the natural environment without fulfilling an intended function (GESAMP 2016; Hartmann et al. 2019). Approximately a third of all the plastics that have been produced is currently in use, and the other two-thirds have been discarded (Geyer, Jenna R. Jambeck, and Kara Lavender Law 2017). While the majority of dis-carded plastics are expected to have been handled appropriately, by either being stored in landfills (21 to 42 %), incinerated (approximately 13%), or recycled (6 to 26%), a signif-icant proportion (>20%) has been mishandled and leaked into the environment (Geyer, Jenna R. Jambeck, and Kara Lavender Law 2017; Alimi, Farner Budarz, et al. 2018). Figure 1.1 (outer circle) presents a breakdown of polymers produced and discarded in 2015 and shows that most of the discarded items come from products with short du-rations of use (<5 years), such as packaging, consumer and institutional products (C&I Products). While most plastic debris is generated from mismanaged plastic waste, plastic debris is also produced incidentally by the aging of plastic materials in the environment. Examples of this latter source are the production of tire wear particles during driving, the release of fibers when washing synthetic textiles, and the aging of fishing nets and ropes during use (Wik and Dave 2009; Zubris and Richards 2005). Finally, a minimal portion of plastic debris comes from intentionally produced plastic particles lost in the environment. These particles are used in agricultural, cosmetic, or medical applications (e.g.: seed coatings, microbeads, and drug delivery). They are introduced directly into the environment or indirectly through wastewaters (Mitrano and Wohlleben 2020). All plastic pieces that come from the incidental degradation of plastics in the environment are defined as secondary plastics, as opposed to primary plastics which are produced intentionally.
All plastic debris is subject to degradation by environmental factors such as sunlight, physical stress, heat, biological activity, etc. (GESAMP 2015; M. Wagner and Lam-bert 2018 and references therein). Although most plastics are designed to be durable, their degradation in the environment is observable on human time scales (Chamas et al. 2020). Rates of plastic degradation and the type of degradation that occur (e.g.: crack-ing, peeling, solubilizing, etc.) depend upon the interplay between plastic composition and environmental conditions (Hahladakis et al. 2018; Min, Cuiﬃ, and Mathers 2020; ter Halle et al. 2016). The most eﬃcient pathways for plastic degradation are photo-oxidation by sunlight irradiation and mechanical abrasion by physical stress (e.g.: by the action of waves and wind) (Chamas et al. 2020; Efimova et al. 2018; Gewert, Plassmann, and MacLeod 2015; Karin Mattsson et al. 2021; Min, Cuiﬃ, and Mathers 2020; Y. K. Song, Hong, Eo, et al. 2020). To a lesser extent, thermo-oxidation, hydrolysis, and biological degradation can also degrade plastic objects (Dawson et al. 2018; Julienne, Delorme, and Lagarde 2019; Min, Cuiﬃ, and Mathers 2020; Zettler, Mincer, and Amaral-Zettler 2013). Plastic degradation is an overarching term that includes many environmental transforma-tions, such as fragmentation and surface oxidation, enzymatic degradation, and leaching of additives (Andrady 2011). All of these transformation processes are interconnected. For example, fragmentation can accelerate the leaching of additives. Surface oxidation can increase particle’s stiﬀness and ability to be fragmented. Ultimately, since plastics are carbon-based materials, they can be mineralized into CO2 and the elementary forms of monomers and additives (Lixin Zhu et al. 2019). However, for complete mineraliza-tion to occur, optimal conditions are required that are rarely met (Chamas et al. 2020; Harrison et al. 2018). Therefore, as illustrated in Figure 1.2, plastic debris persists in the environment as particulate degradation products (i.e.: non-dissolved and non-elemental).
Due to their diversity of sources and degradation pathways, secondary plastic parti-cles can take on an infinite combination of compositions, shapes, surface properties, color, density, hardness, etc. (Kooi, Primpke, et al. 2021; Rochman et al. 2019). The projected increase in secondary particulate plastics and the diverse properties of these contami-nants make them a major environmental concern. In particular, secondary plastics with submicrometric size, which have been called nanoplastics (Gigault, Halle, et al. 2018; Hartmann et al. 2019), are a particular concern since they may form a substantial pro-portion of total stocks of plastic debris but are undetectable with standard instruments and methods (K. Mattsson, L.-A. Hansson, and Cedervall 2015).
Transport and transfer of plastic debris
A material flow analysis (MFA) calculated that approximately 4.8 to 12.7 million metric tons (Mt) of plastic debris generated in coastal countries would enter the oceans in 2010 (J. R. Jambeck et al. 2015). This yearly input of plastics into the ocean is expected to increase in the future if there are no significant changes in the quantity of waste generated and the waste infrastructure (J. R. Jambeck et al. 2015). However, the mass of all plastic debris accumulated at the ocean’s surface since the invention of plastic is orders of magnitude smaller than the estimated yearly input. To evaluate this mass, models have combined ocean circulation models with direct measurements of plastics debris with sizes ranging from 0.33 to 200 mm. According to the highest estimate, there are only 236 000 metric tons of plastics at the ocean surface, representing significantly less than 1% of the plastic expected to be present (van Sebille, Wilcox, et al. 2015). In ocean gyres (zones where oceanic currents converge), where models predict that most plastic debris accumulates, the measured concentration of plastics has been at a steady state in recent years (the 1980s to 2010s)(F. Galgani et al. 2021). Rapid increases in the concentration of plastic debris at the ocean surface have been recorded either i) close to anthropic sources and at the early stages of mass plastic production (Thompson 2004) or ii) in remote areas and decades after the mass production of plastics has started (F. Galgani et al. 2021).
Several hypotheses have been advanced to explain the discrepancy between the plas-tic stocks expected to be at the oceans’ surface and those measured. While none of the hypotheses can on its own explain two orders of magnitude diﬀerence, some hypotheses help the numbers of measured and expected stocks to converge:
Overestimation of the mass of plastic expected at the ocean surface:
Hypothesis 1: The uncertainties in the MFA used to estimate stocks and flows of plastic at the Earth-scale (e.g.: the amount of waste generated, collected through formal and informal methods, and discarded as well as its rate of transfer into oceans) have overestimated inputs of plastic debris into the ocean (van Sebille, Wilcox, et al. 2015; L. Weiss et al. 2021).
Hypothesis 2: The rates at which plastic debris deposits to the shorelines or sink to the ocean floor could have been underestimated, leading to an overestimation of concentrations expected at the ocean surface (van Sebille, Wilcox, et al. 2015)
Underestimation of the mass of plastic measured at the ocean surface:
Hypothesis 3: Plastics larger than 200 mm, which were omitted from models of plastic concentrations at the ocean surface, could constitute a significant portion of the total stocks of plastic (van Sebille, Wilcox, et al. 2015).
Hypothesis 4: The depth at which plastics are located at the ocean surface does not coincide with the depth at which plastics were sampled, leading to an under-estimation of plastic stocks at the ocean surface (Poulain et al. 2019). Indeed, the depth at which small plastic debris (0.3 to 5 mm) is located can vary due to wind turbulence and biofouling (the covering of surfaces by organisms and their exudates, which dynamically modifies their buoyancy) (Kooi, Nes, et al. 2017).
Hypothesis 5: The missing fraction corresponds to small microplastics (1 to 1000 µm) and nanoplastics (<1µm). This smaller plastic debris may have been omitted during sampling since the sampling nets had a minimum mesh size of 330 µm. Even particles ranging from 330 to 1000 µm, which are expected to be captured during sampling, are often lacking in the size distribution of plastic debris(A. Cózar et al. 2014). This is attributed to the analytical challenges involved in measuring small carbon-based particles in natural matrices (Mintenig et al. 2018).
The overestimation of plastic waste input into the oceans (Hypothesis 1), as well as the omission of plastics larger than 200 mm (Hypothesis 3) and plastics located at diﬀer-ent depths (Hypothesis 4) can explain a certain degree of discrepancy between expected stocks and measured stocks. However, these values are expected to be relatively constant over time and therefore cannot explain how plastic debris at the ocean surface has reached a steady state. To explain how plastic concentration at the ocean surface appears to have reached an apparently constant concentration which is lower than predicted concentra-tions, rates of plastic debris removal from the ocean surface by sinking, beaching onto shorelines (Hypothesis 2), and fragmentation down to micrometric and submicrometric sizes (Hypothesis 5) are all preferable hypotheses. In particular, small microplastics (1 to 1000 µm) and nanoplastics (<1 µm) may form an essential fraction of the plastic budget. This hypothesis has been supported by:
– Scarcity of plastics smaller than 1 mm in size distributions of samples (A. Cózar et al. 2014);
– Experimental studies that observed the formation of nanoplastics by abrasion, photo-oxidation, and digestion (Karin Mattsson et al. 2021; Lambert and M. Wag-ner 2016; Gigault, Pedrono, et al. 2016; Dawson et al. 2018);
– Identification of increasingly small microplastics (Primpke, Christiansen, et al. 2020; Primpke, Cross, et al. 2020) and nanoplastic signatures (Ter Halle et al. 2017) with improved analytical methods.
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 : Vs = 2 ρp − ρf grp2 (1.1) 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: D = kB T (1.2) 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).
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
2.2 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
3 Deposition of environmentally relevant nanoplastic models in sand during transport experiments
3.2.1 Dispersions of nanoplastic models
3.2.2 Charge characterization
3.2.3 Size characterization
3.2.4 Transport in porous media
3.3 Results and Discussion
4 Deposition of nanoplastics: The roles of size polydispersity and natural organic matter
4.2 Materials and Methods
4.3 Results and Discussion
5 Micro- and nanoplastics’ transfer in freezing saltwater: Implications for their fate in polar waters
5.2 Materials and Methods
5.3 Results and Discussion
5.5 Supplementary Data
6 Conclusion and Perspectives
List of Figures
List of Tables