Science assessment of plastic pollution

Official title: Science assessment of plastic pollution

Environment and Climate Change Canada

Health Canada

October 2020

Cat. No.: En14-424/2020E-PDF
ISBN 978-0-660-35897-0

List of abbreviations

Executive summary

Plastics are among the most universally used materials in modern society. Since the 1950s, the production and use of plastics has been increasing faster than that of any other material, mostly due to their durability and low cost. However, the improper management of plastic waste has led to plastics becoming ubiquitous in all major compartments of the environment. Plastic that is discarded, disposed of, or abandoned in the environment outside of a managed waste stream is considered plastic pollution.  Plastic pollution has been detected on shorelines, and in surface waters, sediment, groundwater, soil, indoor and outdoor air, food and drinking water. In Canada, it is estimated that 1% of plastic waste enters the environment. In 2016, this amounted to 29 000 tonnes of plastic pollution. Since plastic degrades very slowly and is persistent in the environment, the amount of plastic pollution is anticipated to continue to increase over time. There are growing concerns that plastic pollution may adversely impact the health of the environment and humans.

The purpose of this report is to summarize the current state of the science regarding the potential impacts of plastic pollution on the environment and human health, as well as to guide future research and inform decision-making on plastic pollution in Canada. It provides a review of the available information on plastic pollution, including its sources, occurrence, and fate, as well as on the potential effects of plastic pollution on the environment and human health. This report is not intended to quantify the risks of plastic pollution on the environment or human health, but rather to survey the existing state of science in order to guide future scientific and regulatory activities. 

In an environmental context, plastics are often categorized by their size, with macroplastics being larger than 5 mm and microplastics being less than or equal to 5 mm. Plastic waste can be released into the environment as complete materials (e.g., discarded single-use or short-lived products, such as plastic bags and straws), as large pieces of plastics (e.g., fragments of plastic products) or as microplastics (e.g., microfibres released from washing of clothes or microbeads released through wastewater). Microplastics can also be formed through the breakdown of larger plastic items in the environment.

While plastics can degrade, the rate at which they break down is slow and can be affected by multiple factors, such as temperature and light. In water, the rate of degradation is temperature dependent, being slower in cold water. The lack of exposure to sunlight also slows down the degradation of plastics. While oxidation can promote the degradation of plastics in soil, the rate of degradation is still slow. Although biodegradable, compostable, biobased, and oxo-degradable plastics are increasingly being used as alternatives to conventional plastics, there is a lack of significant evidence that they will fully degrade in natural environments. Further studies would assist in understanding their environmental impact, particularly in comparison to conventional plastics.

Plastic pollution is found in aquatic and terrestrial environments, as well as in indoor and outdoor air, and arises from various sources. For example, plastic may enter the aquatic environment as a result of litter, mismanaged waste, and abandoned, lost or discarded fishing gear, or may be deposited in the terrestrial environment from agricultural activities. Additionally, microplastics removed from wastewater settle in sewage sludge can be released to land through the application of biosolids. Moreover, release of microfibres from wastewater treatment systems is known to represent a source of microplastic pollution. Products available to consumers that are discarded to the environment or not properly managed may also represent a source of plastic pollution. Sources of microplastic pollution to indoor air include the shedding of fibres from clothing, furniture, carpeting and household goods, while microplastics in outdoor air are influenced by various sources including vehicle tire wear and tear.

In Canada as well as internationally, single-use plastics make up the bulk of macroplastics found on shorelines. The most common litter items collected on Canadian shorelines include cigarette butts, bottle caps, plastic bags, plastic bottles, and straws. Microplastic particles such as fragments and pellets are also found on shorelines where they accumulate within the organic matter along the strandline. Generally, a greater abundance of plastic pollution has been found in areas with high human and industrial activity, notably in the Great Lakes. 

Microplastic particles are also found in fresh and marine surface waters. Globally, microfibres are one of the most common types of microplastics found in water. However, it is recognized that there is a lack of standardized, high-quality methods for sampling plastics, particularly for measuring, quantifying and characterizing microplastics.

Microplastics are also found in sediment and soil. Through various mechanisms, such as the formation of biofilms—layers of microorganisms that form on a surface—microplastics in surface waters may eventually sink, leading to the accumulation of microplastics in the sediment of both freshwater and marine environments. Soils are also expected to act as a major sink for plastic particles, as microplastics are likely to remain in soils for long periods due to factors such as vertical transport, which pulls particles down from the surface, slowing their degradation. Currently, only limited evidence is available on the occurrence of microplastics in groundwater, although it has been hypothesized that microplastics may travel from soil into groundwater.

Air is also anticipated to be an important pathway for microplastic transport, and microplastics have been detected in both indoor and outdoor air. While there are no Canadian data available on the occurrence of microplastics in air, limited data from other parts of the world suggest that concentrations may be higher in indoor air than in outdoor air. Indoors, microplastics are also found in settled house dust. 

Current data on the occurrence of microplastics in food are limited, and most available information concerns microplastics found in seafood, specifically fish and shellfish from marine environments. In fish, microplastics have been found in both muscle tissue and the gastrointestinal tract, mostly as fragments and fibres. Microplastics have also been detected in mussels, clams, oysters, scallops and snails, and in a very small number of other foods, such as salt. 

Internationally, a limited number of studies have investigated the presence of microplastics in tap and bottled water. Of these studies, few are considered reliable due to concerns with quality assurance measures. While the available studies indicate that microplastics have been detected in bottled water samples purchased from outside of Canada, the concentrations do not correlate with bottle type (i.e., plastic, glass or beverage carton) and vary depending on the use conditions (i.e., single-use versus multi-use bottles). The sources of microplastics in bottled water are still unknown and further research is required. In the case of tap water, some studies have detected microplastics while others have not. It is anticipated that drinking water treatment will remove a large proportion of microplastic particles.

Plastic pollution has been shown to impact organisms and their habitats. Macroplastics have been demonstrated to cause physical harm to environmental receptors on an individual level and to have the potential to adversely affect habitat integrity. Physical harm to biota is often a result of entanglement or ingestion. Entanglement can lead to suffocation, strangulation, or smothering, and a high frequency of reported entanglements have led to direct harm or mortality. Ingestion can lead to direct harm through physical damage; it can block airways or intestinal systems leading to suffocation or starvation. An increasing amount of plastic pollution in water bodies may also adversely affect ecosystem function, biodiversity, and habitat integrity. Plastics can act as transport mediums for organisms, microorganisms, or other organic matter, which can alter ecosystem dynamics. 

The observed effects of microplastics on biota are primarily driven by physical effects. Published studies on the ecotoxicological effects of microplastics report conflicting observations, even for the same endpoint in the same species. This conflicting information could be addressed by developing and using standard approaches for testing the effects of microplastics on environmental organisms, using environmentally relevant testing materials, and developing an understanding of the impact of shape, size and chemical composition on ecotoxicological effects.

Humans may be exposed to microplastics via the ingestion of food, bottled water, and tap water, as well as through the inhalation of indoor and outdoor air. However, information on the human health effects of microplastics is limited, and further research is required to better inform target tissues, threshold doses, and mode of action. Some associations between exposures to high levels of microplastics and adverse health effects in laboratory animals and in humans have been reported, but these health effects cannot be linked to exposure in the general population. Occupational inhalation exposure studies show associations between work in microplastic-related industries and increased incidence of various respiratory symptoms and diseases. Conflicting observations have been made for cancers of the respiratory tract and digestive system. 

Effects observed in animal studies are primarily associated with tissues related to where particles enter the body (e.g., effects on the digestive system after oral exposure and on the respiratory tract after inhalation). Effects following oral exposure include inflammation of the liver, oxidative stress, metabolic changes, and altered gut microbiota. Movement of a small fraction of microplastic particles to lymphatic tissues has also been observed. Although the current scientific literature does not identify a concern for human health, there are insufficient data to allow for a robust evaluation of the potential human health risks of ingested microplastics.

Effects in the respiratory tract are likely related to the physical impact of microplastics as particulate matter and include oxidative stress, cytotoxicity, inflammation, and development of foreign body granulomas. In inhalation studies, movement of a small fraction of microplastic particles to lymphatic or systemic tissues has also been observed. No dose-response relationship has been observed in mortality, survival time, behaviour, clinical observations, or tumour incidence from inhalation exposures. 

In addition to physical impacts, there are concerns that plastics may serve as a means of transport for other chemicals. Since plastics can contain unbound monomers and chemical additives and can sorb persistent organic pollutants from the environment, it is possible that these substances may be transported to organisms or humans, where they may then be released. The extent of release is expected to depend on a variety of factors, such as the properties of the receiving environment, the plastic particle, and the bound chemical. The current literature suggests that, while the transport of chemicals via plastics is possible, the impact to biota is likely limited, and recent international reviews indicate that there is likely a low health concern for human exposure to chemicals from ingestion of microplastics from food or drinking water (EFSA 2016; FAO 2017; WHO 2019). Many of the chemicals observed to be bound to plastic particles have been assessed by various programs at Environment and Climate Change Canada and Health Canada.

Plastics can also provide a habitat for microorganisms, including potential pathogens, through the formation of biofilms. There is currently no indication that microplastic-associated biofilms would impact human health. In addition, despite very limited data, it is anticipated that drinking water treatment would inactivate biofilm-associated microorganisms. 

In order to advance the understanding of the impacts of plastic pollution on the environment and human health, it is recommended that research be carried out in the following areas to address the key knowledge gaps identified in this report:

• Developing standardized methods for sampling, quantifying, characterizing, and evaluating the effects of macroplastics and microplastics;
• Furthering the understanding of human exposure to microplastics;
• Furthering the understanding of the ecotoxicological effects of microplastics;
• Furthering the understanding of the effects of microplastics on human health; and
• Expanding and developing consistent monitoring efforts to include poorly characterized environmental compartments such as soil.

Given the increasing amounts of plastic pollution in the environment and the demonstrated ability of macroplastics to harm biota, it is anticipated that the frequency of occurrence of physical effects on individual environmental receptors will continue to increase if current trends continue without mitigation measures. 

In accordance with the precautionary principle, action is needed to reduce macroplastics and microplastics that end up in the environment.

1. Introduction

Plastics are part of the everyday lives of Canadians and populations around the world. Plastics are low cost, durable materials and can be used in a variety of applications (CCME 2018). For these reasons, global plastic production has been increasing over the past several decades at a rate faster than that of any other material (Geyer et al. 2017; CCME 2018). In Canada, total sales of plastic were estimated to be $35 billion in 2017, with approximately 4 667 kt introduced to the Canadian market in 2016 (ECCC 2019a). Plastics are used in a variety of industrial sectors, and demand for plastic products continues to grow.  

Of the 4 667 kt of plastics that entered the Canadian market in 2016, an estimated 3 268 kt were discarded as waste (ECCC 2019a). Of that plastic waste, an estimated 29 kt (or 1%) were discarded outside of the normal waste stream (i.e., not landfilled, recycled or incinerated) in 2016, through direct release to the environment or through dumps or leaks. An estimated 9% of the remaining plastic waste was recycled, 86% was landfilled, and 4% was incinerated for energy recovery (ECCC 2019a). 

In a global context, Geyer et al. (2017) have estimated that only 30% (2 500 000 kt) of all plastics ever produced are still in use. This means that 6 300 000 kt of global cumulative plastic waste was created between 1950 and 2015. As illustrated in Figure 1, if plastic manufacturing continues at its current pace, the accumulation of plastics will continue to accelerate. It is estimated that by 2050, 12 000 000 kt of plastic waste will have been discarded globally to landfills or the environment (Geyer et al. 2017).

With the growing public and scientific concern about the ubiquity of plastic pollution, there has been increasing global media attention on the potential impacts of plastic pollution on human health and the environment (CCME 2018; ECCC 2019b; SAPEA 2019). The Government of Canada has put forward Canada’s Plastics Science Agenda (CaPSA), which aims to align current and future research investments across a range of disciplines (ECCC 2019b). The CaPSA framework identifies several key research priorities, including the detection of plastics in the environment, understanding and mitigating potential impacts on wildlife, human health and the environment, plastic design and alternatives, sustainable plastic production, and recycling and recovery.

1.1 Scope

This report summarizes the current state of the science on the potential impacts of plastic pollution on the environment and human health and informs future research and decision-making on plastic pollution in Canada. For the purposes of this report, plastic pollution has been divided into two main types: macroplastics (plastics greater than 5 mm in size) and microplastics (plastics less than or equal to 5 mm in size). This report discusses the sources, occurrence, and fate of plastic pollution in the environment, as well as the potential impacts of plastics on human health and the environment. In the draft science assessment, published in January 2020, information identified up to June 2019 was considered for inclusion, in addition to the August 2019 World Health Organization (WHO) report on microplastics in drinking water and the October 2019 Ocean Wise report on microfibres (Vassilenko et al. 2019). In finalizing the science assessment, a review of information published up to March 2020 was conducted (see Appendix A and Section 5.1.1 for relevant new information), in addition to considering comments submitted through public consultation. 

This report is a review of the current state of the science on plastic pollution. It is not intended as a substitute for chemical risk assessment, and it is similar to the approach taken for the Science Summary on Microbeads (ECCC 2015). Typically, a chemical risk assessment is conducted to assess the potential for risk to the environment and human health associated with a substance. However, significant data gaps currently exist that preclude the ability to conduct a quantitative risk assessment, including a lack of standardized methods for monitoring microplastics and characterizing the environmental and human health effects of plastic pollution, as well as inconsistencies in the reporting of occurrence and effects data in the scientific literature (Gouin et al. 2019). Indeed, risk assessment frameworks for evaluating the potential risks associated with plastic pollution are currently under development. For example, see Gouin et al. (2019) for a discussion on the development of an environmental risk assessment framework for microplastics. 

As the focus of this report is on plastic pollution, it is limited to a review of the occurrence of macroplastics and microplastics resulting from plastic waste entering the environment and does not examine non-environmental sources (e.g., via direct exposure from products available to consumers or self-care products^{Footnote 1} ). Moreover, it does not review the economics of waste management practices or evaluate the efficacy of waste management streams (e.g., recycling). 

1.2 Terminology

This report discusses plastic pollution in an environmental context. In this context, plastics are often categorized by their size. The term microplastic was originally used to differentiate between substances that could only be visualized through a microscope and larger macroplastics (ECCC 2015). However, there is no single definition of what constitutes a microplastic. For the purpose of this report, plastic particles less than or equal to 5 mm in size are defined as microplastics, while plastics greater than 5 mm are defined as macroplastics. Microplastics can be further defined as primary or secondary microplastics. Primary microplastics are intentionally produced plastic particles (such as pellets, powders, and beads) that are either intended for use as microplastics or as precursors for the production of plastic or plastic-containing products. Primary microplastics are widely used as abrasives in a variety of applications (UNEP 2016). Secondary microplastics are not produced intentionally, but are the result of the breakdown and fragmentation of larger plastic items (SAPEA 2019). Furthermore, microfibres are a specific type of secondary microplastic defined as being fibrous in shape and less than or equal to 5 mm in length. Nanoplastics are considered to be a subset of microplastics. They are primary or secondary microplastics that range from 1 to 100 nm in size in at least one dimension. Nanoplastics occur largely as a result of secondary sources of plastic pollution (i.e., the breakdown of larger plastics) (Rist and Hartmann 2018). This report will focus on plastics greater than 100 nm in size (i.e., microplastics and macroplastics).

In this report, plastic waste is considered to be plastics that enter the waste stream (e.g., landfilled, recycled or incinerated), whereas plastic pollution is considered to be plastic that is discarded, disposed of, or abandoned in the environment outside of a managed waste stream. In the scientific literature, plastic pollution has been referred to by a number of terms, such as plastic debris or plastic litter. This report will use the terms plastic pollution or plastic pollutants. Furthermore, in this report the term litter refers to any persistent, manufactured, or processed solid material discarded, disposed of, lost, or abandoned in the environment, including plastics, textiles, glass, metal, ceramics, and other persistent synthetic materials. This term will be used when the proportion of plastic pollution reported in the literature is unclear.

2. Composition, properties and uses

All plastic materials are formed from long-chain polymers of very high molecular weight, often measured in the hundreds of thousands of kilodaltons (Sperling 2006). Synthetic polymers first appeared in the early 20th century, leading to the manufacture of plastic products such as Bakelite and nylon (commercial name for polyamides). Since then, polymer science has evolved, with a greater mechanistic understanding of the interrelationships between polymer structure, morphology, and physical and mechanical behaviour. This has resulted in the production of a myriad of plastic materials with varying physical and chemical properties. 

Polymerization, the synthesis of polymers, can occur following one of two main processes: chain polymerization or stepwise polymerization. The process used to form polymers greatly influences their physical properties. Common chain polymer structures include polyethylene (PE), polypropylene (PP), polystyrene (PS), and polyvinyl chloride (PVC), whereas common stepwise polymers include nylons, polyethylene terephthalate (PET), polycarbonate (PC), and polyurethane (PU). These represent many of the most common forms of plastics typically found as environmental contaminants (Sperling 2006). 

The physical properties of plastic, such as rigidity, flexibility, and elasticity, are influenced by the polymer’s molecular weight distribution and organization of polymer chains (Sperling 2006; Verschoor 2015). Generally, high molecular-weight polymers with a complex organization that leads to strong covalent bonds between the polymers can result in the formation of a rigid plastic with a high melting point. In contrast, linear polymer organization with low molecular-weight distribution results in a more flexible plastic with a lower melting point. Combinations of different molecular weight distributions, different polymer chain organization, and/or blends of different types of polymers can produce a material that will be effective for its intended use (Sperling 2006). 

Furthermore, many polymers are subject to additional processes aimed at enhancing efficacy with respect to an intended functionality. For instance, when heated, a linear polymer will flow, resulting in the formation of a thermoplastic (Sperling 2006). Thermoplastics are polymers commonly found in plastics that can be melted and reshaped into new objects. Commonly used thermoplastics include PVC, PE, PS, and PC (ECCC 2019c). To prevent flow upon heating, polymers can be cross-linked to produce a thermoset plastic (Sperling 2006). Thermoset plastics are polymers that are used for their resistance to mechanical forces, chemicals, wear, and heat, but they cannot be re-melted to form new objects. Examples include PU and unsaturated polyester polymers (ECCC 2019c).

Chemical additives can be added to polymers during production to alter the properties of plastics (Rochman et al. 2019). There are several categories of additives, including stabilizers and functional agents. Polymer stabilizers maintain the inherent properties of the material by protecting it against oxidative degradation. They include substances such as anti-oxidants, light stabilizers, metal deactivators, and ultra-violet absorbers. Functional agents can enhance the mechanical strength of a polymer or impart new characteristics. Examples of functional agents include flame-retardants, anti-static agents, lubricants, and plasticizers (ECCC 2019c). For instance, plasticizers can be added to soften a polymer by lowering its glass transition temperature or reducing the degree of crystallinity or melting point (Sperling 2006).

Polymer production can begin with either the use of recycled or recovered plastics or with natural resources (i.e., petroleum or plant-based starting material). These polymers are then used to manufacture plastic products (ECCC 2019c). Although many different types of plastic polymers are used in Canada, domestic plastic production is dominated by five polymer types. PE accounts for the majority of plastic production, with 3 700 kt produced in 2017, followed by PVC (210 kt), PET (166 kt), PU (122 kt) and polyamides (PA) (116 kt) (ECCC 2019c). Of the 4 800 kt of plastic polymers produced in Canada in 2016, 77% was exported. Further, there is a domestic demand of 3 800 kt, 71% of which is fulfilled through imports (ECCC 2019a).

The majority of plastic products in Canada are found in the packaging and construction sectors. Other major sectors include the automotive, electronic and electrical equipment, textiles, and agriculture sectors (ECCC 2019c). Examples of applications of various polymers are presented in Table 2‑1.

^a Personal communication, email from the Transportation and Manufacturing Division, National Research Council Canada, to the Ecological Assessment Division, Environment and Climate Change Canada, dated August 15, 2019; unreferenced

The packaging sector is the largest user of plastics in Canada, accounting for 33% of end-use plastics introduced to the market in 2016. Examples of plastic packaging products include plastic bags, water and soft drink bottles, as well as various packaging used for pharmaceuticals, toiletries, and cleaning compounds. PE is very commonly used in packaging, specifically for films and flexible packaging. The main types of PE are low-density polyethylene (LDPE), linear low-density polyethylene (LLDPE), and high-density polyethylene (HDPE). These types of PE form the majority of PE production globally. Some examples of LDPE/LLDPE applications are squeeze bottles, toys, carrier bags, and general packaging.  LDPE is generally used in heavier duty films, such as high durability bags and protective sheeting, due to its toughness, flexibility, and relative transparency. HDPE possesses good chemical resistance and is used for packaging many household and industrial chemicals such as detergents and bleach. It is also used in thin-gauge carrier bags, chemical drums, toys, food wrapping material, and kitchenware. In addition to PE, other plastic polymers are used in the packaging sector, such as PVC, PET and PP (ECCC 2019c). 

Construction is the second-largest end-use market for plastics in Canada, accounting for approximately 26% of all end-use plastics generated in 2016 (ECCC 2019a). Primary uses of plastics in the construction sector include plastic and foam building and construction materials, paints and coatings, profile shapes, and reconstituted wood and plywood. Plastics are broadly used in the construction of all types of buildings and are especially used in thermal insulation materials, as well as waterproofing and sealant materials. PVC is widely used in siding and window applications, floor and wall covering products, as well as pipe and pipe fittings. Clear PC sheets are used as a substitute for glass in greenhouses, transit shelters, and covered walkways due to its resistance to weathering. PU foam is used as insulation in commercial and residential properties (ECCC 2019c). 

In an effort to improve fuel efficiency through weight reduction, the automotive sector has increased its use of plastics. While many different types of plastics are used in the sector, PU, PP, and PVC make up the vast majority of total plastics used in a vehicle. PU is used in cushioning applications such as seating, PP is used in automobile interiors, and PVC is used for faux leather. PC can be used to replace glass in cars, while foam, plastic, and fibre composites can be used in door panels, dashboards, and hoods (ECCC 2019c). In this report, vehicle tires are considered a source of plastic pollution via the release of tire wear particles. Yet, it is recognized that whether or not rubber is considered a type of plastic is the subject of some debate at this time, and while some scientific publications consider rubber to be plastic, others do not.

Other end-use sectors include the electronic and electrical equipment, textile, and agriculture sectors. Plastics are used in the electronics sector for computer and phone parts, as well as items such as electrical wires and cables. The textile sector uses plastics for fibres in carpets, rugs, mats, and clothing. In the agriculture sector, plastics are used for fertilizer and pesticide packaging (e.g., agricultural films, mulches, and greenhouses) (Ekebafe et al. 2011; ECCC 2019c).

Given the variety of plastic materials that can be produced, the physical and chemical properties of plastic particles present in the environment will be complex (Rochman et al. 2019). With respect to shape and size, primary microplastics are intentionally engineered to be a particular size (e.g., virgin resin pellets used in plastic manufacturing processes) and will therefore likely show less variation than secondary microplastics. Secondary microplastics can have a range of shapes, including spheres and cylinders, but also fragments, fibres, and films (Kooi and Koelmans 2019). Secondary microplastics are also highly variable in size and density. Recognizing the inherent challenge associated with defining the physical properties of microplastic particles observed in the environment, Kooi and Koelmans (2019) suggest a method aimed at defining and characterizing the distributions of properties most commonly encountered. The approach proposed by Kooi and Koelmans (2019) may prove useful in developing tools for monitoring plastics in the environment, providing a greater mechanistic understanding of the environmental fate of microplastics, and allowing for easy comparison between studies. 

Microplastics can exist as fibres, fragments, spheres, pellets, films, and foams, as shown in Figure 2. In general, certain shapes of microplastics originate from certain plastic products. For example, fibres are typically shed from fabrics, such as clothing and upholstery, whereas pellets are typically from industrial feedstock (Rochman et al. 2019). 

Density is a key property that influences the environmental fate of plastics (Rochman et al. 2019). Densities of plastic polymers such as PE, PS and PVC can range from 0.9 to 2.3 g/cm³ (WHO 2019). Table 2-2 presents the densities of various plastic polymers. Polymers with a density greater than 1 are denser than water and are expected to sink, while those with a density less than 1 are expected to float. By analogy, the environmental fate and transport of macroplastics or microplastics released to the atmosphere are also likely to be influenced by their density. For example, denser microplastics are less likely to be readily dispersed by the wind (Rochman et al. 2019). The density of plastics and their buoyancy in water can also be influenced by the coating of plastics with microorganisms, algae, or plants (i.e., biofilms) (Woodall et al. 2014). Other factors, such as shape and size, can also govern the fate of plastics in the environment (Rochman et al. 2019).

^a Hidalgo-Ruz et al. 2012

3. Sources of plastic waste and pollution

3.1 Sources of plastic waste

In Canada, the main industrial sectors contributing to the estimated 3 268 kt of plastic waste discarded in 2016 are presented in Table 3-1. Plastic packaging is the single largest contributor of plastic waste, followed by the automotive, textile and electrical and electronic equipment sectors. In 2016, 33% of the plastics entering the Canadian marketplace was for use in packaging. Based on international data, it is estimated that 40% of all plastic production is used for packaging, a significant portion of which is used for the food and drink sector (UNEP 2016). Due to the extremely short lifecycle of plastics from packaging (i.e., most plastic packaging is single-use in nature) compared to plastics from other sectors, packaging accounted for 47% of the plastics discarded in Canada in 2016. Plastics generated from other industrial sectors, such as the automotive (vehicle parts and components, excluding tire wear) and construction sectors, have longer lifecycles and therefore represent a smaller proportion of annual plastic waste as compared to packaging, which is typically discarded shortly after use (ECCC 2019a).

^a Films (including plastic bags), bottles and other items from sectors including food and beverage, healthcare, consumer packaged goods, and cosmetics and personal care, among countless other applications.
^b Includes chemical products, toys, household furniture, etc. See ECCC (2019a) for a complete description.

3.2 Sources of plastic pollution

The sources of global plastic pollution are varied, and actual quantities of plastic pollution are largely unknown (UNEP 2016). Plastics that are discarded to the environment or not properly managed represent sources of plastic pollution. Land-based sources of macroplastics to the marine environment include packaging, construction materials, household goods, and items related to coastal tourism (UNEP 2016). Land-based sources of microplastics around the world include cosmetics and personal care products, synthetic textiles and clothing, terrestrial transport (i.e., tire wear), and plastic producers and fabricators (i.e., accidental loss of resin pellets) (UNEP 2016; SAPEA 2019). With respect to sea-based sources of plastic pollution, the fisheries, aquaculture, and shipping sectors are major contributors (GESAMP 2016; UNEP 2016). Plastics in these sectors may be lost at sea by accident, abandonment, or deliberate disposal (UNEP 2016; SAPEA 2019). Macroplastics and microplastics from land- and sea-based sources can enter the ocean through various entry points (e.g., wastewater, rivers, coasts), depending on the region (UNEP 2016).

Products available to consumers discarded to the environment or not properly managed may also represent a source of plastic pollution in the environment. While knowledge of the source of primary microplastics (i.e., the type and amount of microplastics intentionally used in products available to consumers) in Canada is limited, secondary microplastics may arise from the breakdown and fragmentation of macroplastics released to the environment. This may include items such as toys, plastic gloves, appliances, electronics, mattress covers and flooring, as well as plastic materials used in packaging (Table 3‑1).   

3.2.1 Sources to water

Plastic pollution in the aquatic environment can arise from plastics released during land-based activities (e.g., through littering, inadequate waste management, landfill leachate, the use of plastics in agriculture, land application of biosolids, or direct release following abrasion or maintenance of plastic products (Boucher and Friot 2017; Alimi et al. 2018), from the deposition of airborne microplastics onto water (Hendrickson et al. 2018), from runoff and stormwater (Grbić et al. 2020)), or from water-based sources (e.g., fishing-related litter (Driedger et al. 2015)). Plastic pollution in water may also arise from the accidental release of raw plastic materials, such as spillage during transport (Driedger et al. 2015) and from releases from wastewater effluent (Murphy et al. 2016; Boucher and Friot 2017; Kay et al. 2018).

The Arctic Council’s Protection of the Arctic Marine Environment (PAME) Working Group recently released the Desktop Study on Marine Litter including Microplastics in the Arctic as part of the first phase of a Marine Litter Project. The major sectors highlighted as sources of marine litter in the Arctic were fisheries, aquaculture, shipping, cruise tourism, and offshore resource exploration and exploitation. It is estimated that approximately 640 kt of abandoned, lost or discarded fishing gear are released to marine waters globally each year, accounting for 10% of all marine litter. In addition, releases from communities that are not connected to large waste management systems have been flagged as sources of marine litter (PAME 2019).

Wastewater treatment

When wastewater containing plastics from domestic, commercial, and industrial sources passes through wastewater treatment systemsFootnote 2 (WWTSs), most of the plastics are removed prior to discharge to the aquatic environment.

Based on a review of several published studies, Sun et al. (2019) reported significant reductions in microplastic concentrations when comparing influent and effluent in various WWTSs: concentrations ranged from 1 to 10 044 particles/L for influent and from 0 to 447 particles/L for effluent. While large variations in microplastic concentrations can be observed between WWTSs, this may be due to differences in sample collection and analysis methods, as there are currently no standardized methods for the detection and quantification of microplastics in water. Other factors, such as catchment size, population served, wastewater source (residential, commercial, or industrial), and treatment technology, may also contribute to variations in influent and effluent concentrations and treatment efficiencies (Sun et al. 2019).

According to available data on the microplastic removal efficiencies of WWTSs, standard wastewater treatment systems using primary and secondary treatment processes can effectively remove most microplastics from the effluent before it is released to receiving waters (WHO 2019). Sun et al. (2019) estimated that 50% to 98% of microplastics can be removed during primary treatment, which involves skimming processes and settling stages, with larger particles being preferentially removed. Secondary treatment, which typically involves biological treatment to remove organic compounds, can increase microplastic removal to approximately 86% to 99.8% of microplastics (Sun et al. 2019; Raju et al. 2018). The addition of tertiary treatment can lead to the removal of 98% to 99.8% of microplastics, but removal efficiency is dependent on the type of treatment technology used (Sun et al. 2019). Advanced technologies such as rapid-sand filters, membrane bioreactors, and dissolved-air flotation can remove 95% to 99.9% of microplastics greater than 20 µm (Lares et al. 2018; Talvitie et al. 2017). Mintenig et al. (2019) observed complete removal of microplastics greater than 500 µm and 95% of microplastics less than 500 µm using tertiary filtration.

Given the large volumes of effluent water leaving a WWTS, even a small fraction of microplastics remaining in the effluent water after treatment can translate into high absolute numbers of particles being released to the environment. Effluent discharges have therefore been identified as an important pathway for the entry of microplastics into freshwater sources (Murphy et al. 2016). 

It is estimated that a single WWTS discharges an average of 2 million microplastic particles per day (Sun et al. 2019). In a study conducted at a WWTS near Vancouver, it was estimated that 32 million to 97 million microplastics per day are discharged in effluent (Gies et al. 2018), with fibres and fragments being the most abundant microplastic in the effluent. The study also estimated that of the 1.76 trillion microplastic particles that enter the WWTS each year, 1.28 trillion settle into primary sludge, 360 billion exit in secondary sludge, and 30 billion pass into the secondary treatment effluent and are released into the environment, corresponding to up to 99% removal of microplastics in the WWTS.

The most frequent polymers in WWTS influent and effluent are polyester, PE, PET and PA, with fibres accounting for approximately 52.7% of the microplastics found in wastewater, which is likely attributable to the large amount of fibres released during domestic laundering (Sun et al. 2019). A study conducted by the Swedish Environmental Research Institute found that microfibres were the predominant type of microplastics found in sewage sludge from WWTSs, which is consistent with observations in other studies (Magnusson and Norén 2014; Mahon et al. 2017; Li et al. 2018a).

Microfibres from laundering of textiles also represent a significant source to waterbodies. A report by Ocean Wise detailed the results of a study in which 38 different textile samples were tested for their shedding properties using a custom-designed washing machine test facility. The extent of microfibre shedding varied with the type of textile, with polyester, wool and cotton textiles releasing the largest amounts of microfibres. The report also estimates that the average Canadian household releases 533 million microfibres from laundry every year and that an estimated 878 tonnes of microfibres are released to water following wastewater treatment in Canada and the United States annually (Vassilenko et al. 2019). 

Synthetic textiles and clothing are a large source of microplastic pollution (SAPEA 2019). Microfibres can be released from synthetic fabrics during wear and laundering, as well as from sources such as fishing gear (e.g., fishing nets) (ECCC 2019d). Carney Almroth et al. (2018) and De Falco et al. (2018) counted the number of microfibres released from different types of fabric under different laundering conditions. Both studies found that the use of a detergent increases the number of fibres released during washing. Powdered detergents, which often contain insoluble compounds that are able to create friction with the fabric, enable an even greater number of fibres to be released (De Falco et al. 2018). It has also been noted that powdered detergents have a higher pH compared to liquid detergents. While this is effective for soil removal, it can damage polyester fabrics by way of slow surface hydrolysis (Bishop 1995). Furthermore, exposure of fabrics to chemical detergents can cause the breakdown of synthetic fibres into smaller fibres (SAPEA 2019). The studies found that fleece garments and tightly knit fabrics released the greatest number of fibres during washing. It was found that on average, an adult-sized PET fleece garment releases an estimated 110 000 fibres during washing (Carney Almroth et al. 2018). A wash load of 5 kg of polyester garments was found to release 6 000 000 to 17 700 000 fibres, for an approximate weight of 0.43 to 1.27 g (De Falco et al. 2018).

3.2.2 Sources to soil

Plastic pollution can enter terrestrial environments through various sources including litter, plastic products used in agriculture, such as plastic seed casings, ground covers, and crop mulch. Sources of microplastics to soil include land application of biosolids, plastic pollution, and poorly managed landfills (Alimi et al. 2018). 

The settling stages of the wastewater treatment process result in the production of sewage sludge that contains large amounts of microplastics (Mahon et al. 2017). It is estimated that 99% of microplastics are removed from the influent but are retained in sewage sludge (Magnusson and Norén 2014) and that the properties of microplastics, such as their hydrophobicity and surface charge, can affect their accumulation in the solid phase (Murphy et al. 2016). However, the configurations of WWTSs differ, and thus the removal of microplastics from the influent vary from study to study (Novotna et al. 2019). Microplastics can therefore enter terrestrial environments through the application and use of sewage sludge as fertilizers for agriculture or landscaping purposes (Raju et al. 2018). In Europe and North America, around 50% of sewage sludge is recycled for use as fertilizer, and it is estimated that 44 to 300 kt of microplastics are added to farmlands in North America annually (Nizzetto et al. 2016).

3.2.3 Sources to air

Road traffic-related releases of particles from tire wear and tear are a source of microplastics to outdoor air (Panko et al. 2013, 2019; Kole et al. 2017; Prata 2018). Additional sources of microplastics in outdoor air are thought to include airplane tires, artificial turf, thermoplastic road markings, waste incineration, construction, landfills, industrial emissions, and tumble dryer exhaust, although their relative contributions have not been well established (Dris et al. 2016; Magnusson et al. 2016; Kole et al. 2017; Prata 2018). Deposition and dispersion of all airborne plastic particles from the air may result in accumulations of microplastics in water. Current estimates of the contribution of airborne tire wear and tear particles to water bodies and oceans are varied (e.g., Kole et al. 2017; Sieber et al. 2020; Unice et al. 2019a, 2019b) and additional research is necessary. However, findings suggest that tire wear particles that occur as road dust (i.e., particles that settle rapidly and that are less prone to air dispersion) are a more important contributor to total microplastic pollution in oceans than those found in ambient air.

The primary source of microplastic particles in indoor air is thought to be the shedding of polymeric textile fibres from clothing, furniture, carpeting, and household goods due to wear and tear or abrasion (Sundt et al. 2014; Dris et al. 2016). For example, washing clothing made from synthetic materials has been shown to release microplastics into wastewater, and it is hypothesized that air- or tumble-drying these garments would also cause fragments to be transferred to indoor air, household dust or dryer lint (Wright and Kelly 2017; Prata 2018). Synthetic textile fibres have also been retrieved from a variety of surfaces, including outdoor surfaces, suggesting that clothing and other fabrics may be additional sources of microplastics in both outdoor and indoor air (Rauert et al. 2014; Dris et al. 2016; Prata 2018). 

4. Environmental fate

This section reviews the available data on the fate of macroplastics and microplastics in three environmental compartments: water, soil, and air. It then discusses the persistence of plastics in the environment and the conditions under which they will break down (e.g., transition from macroplastics to microplastics). The fate of biodegradable plastics and biobased plastics is also addressed.

The transport of plastic pollution often follows hydrological pathways (Windsor et al. 2019), with rivers being a key transport pathway (see Figure 3) (Alimi et al. 2018). From rivers, it is expected that the majority of plastic pollution will eventually be transported to the ocean. The mechanisms of transport are poorly understood, but are thought to be influenced by the shape, density, size, and surface condition (i.e., degree of weathering) of the plastic particle. It is also thought that the behaviour of macroplastics differs from microplastics since more energy would be required to transport larger plastics through an ecosystem even if the same transport mechanism is used (Windsor et al. 2019).

4.1 Degradation

Plastic degradation in the environment is slow and can be affected by multiple factors (Andrady 2015; Gewert et al. 2015). Plastics that are exposed to sunlight, oxidants, and physical stress over time will weather and degrade, although the extent of degradation depends on both the environment and the chemical composition of the plastic (Eubeler et al. 2010).

Owing to their chemical structure, common synthetic polymers are durable and can be resistant to degradation. However, there are multiple processes that can bring about the degradation of polymers. These include solar UV-induced photodegradation, thermo-oxidation, hydrolysis, and biodegradation (i.e., degradation by microorganisms). The most common polymers in the environment, such as PE, PP, PS and PVC, possess a carbon backbone that is resistant to biodegradation. Therefore, in order for biodegradation of these polymers to occur, an abiotic degradation step is needed to first break them down into smaller, lower molecular weight fragments (Gewert et al. 2015; Ng et al. 2018). Given that plastic degradation occurs primarily through exposure to sunlight, degradation is most intensive in environments such as the sea surface and beaches (Andrady 2015).

The first visual effects of polymer degradation are changes in colour and cracking of the surface (Gewert et al. 2015). Surface cracking causes the inside of the plastic material to be further available for degradation, eventually leading to embrittlement and physical breakdown upon exposure to abrasive or mechanical forces, such as wind, waves, and physical impacts (Gewert et al. 2015; Ng et al. 2018; SAPEA 2019). Over time, fragmentation can result in plastics no longer being visible in the environment (Selke et al. 2015).

It is important to note that several degradation pathways may take place simultaneously since various factors initiate degradation. For that reason, degradation products may be more diverse than those expected for any specific pathway. In the marine environment, most plastics degrade first at the polymer surface that is exposed and available for chemical or enzymatic attack. Microplastics have higher surface-to-volume ratios than macroplastics and therefore degrade faster than macroplastics, but the process is still slow (Andrady 2015). 

There are numerous gaps in research on plastic degradation. To estimate biodegradation, many studies examine factors such as weight loss, decrease in tensile strength, visual disappearance, or the growth of different microorganisms (Zumstein et al. 2019). 

Plastics often include additives that, when released, may degrade to form other chemicals. In addition, additives such as stabilizers may enhance resistance to degradation. A study by Selke et al. (2015) evaluated the effect of biodegradation-promoting additives on the biodegradation of PE and PET in compost, landfill, and soil environments. They found that none of the additives significantly increased biodegradation in any of the conditions, and there was no evidence that these additives promoted or enhanced biodegradation of PE or PET polymers (Selke et al. 2015).

4.1.1 Biodegradable, compostable, biobased and oxo-degradable plastics

Biodegradable, compostable, biobased and oxo-degradable plastics are often regarded as potential solutions to the accumulation of plastic litter and waste (European Commission 2019; Napper and Thompson 2019). Some of these terms are explicitly defined elsewhere in the context of various certifications (e.g., ASTM D6400, ASTM D6868-19, and ASTM D883-20a). The following provides a brief overview of these plastics as they relate to the issue of plastic pollution.

Biodegradable plastics are a type of plastic that possess heteroatoms along their backbone that render them more susceptible to hydrolytic or enzymatic reactions (Ng et al. 2018). These processes cause the structure to break down into lower molecular weight fragments that microbial cells can assimilate and subsequently mineralize^{Footnote 3} either aerobically or anaerobically. The conditions for biodegradable plastics to break down vary and there is a need to differentiate degradation pathways under different conditions (Lambert and Wagner 2017). For example, some types of biodegradable plastics will not mineralize unless they are exposed to temperatures above 50°C for long periods of time, conditions that are rarely found in the natural environment, but rather in industrial composting facilities (UNEP 2015). 

Compostable plastics are a type of biodegradable plastic that are designed to biodegrade in a managed composting process through the action of naturally occurring microorganisms, typically within a specified time frame (Napper and Thompson 2019). While composting of these products has been explored in Canada, very little post-consumer plastic is managed through industrial composting facilities (ECCC 2019e). Difficulties distinguishing compostable from non-compostable plastics can also create contamination problems for processors. As a result, some certified compostable plastics are not accepted by many composting facilities in Canada (ECCC 2019e).

Biobased plastics are plastics that are synthesized from biomass or renewable resources. Many do not necessarily biodegrade more readily than conventional plastics (European Commission 2019), and unless demonstrated through a complete lifecycle analysis, they do not necessarily confer any superiority to petroleum-based plastics with respect to environmental factors (Vert et al. 2012).

Oxo-degradable plastics, which are sometimes referred to as oxo-biodegradable plastics (UNEP 2015), are formulated using conventional polymers with the addition of heat and UV-activated additives to accelerate their fragmentation into microplastics. While it is expected that accelerated fragmentation would also accelerate degradation, the degree and speed of fragmentation are dependent on environmental conditions such as temperature and light intensity, which vary from day to day, and according to local conditions. Therefore, there is no conclusive evidence that accelerating fragmentation will enable degradation. Given that fragmentation of oxo-degradable plastic requires oxygen and that the majority of plastics in landfills will not have direct access to oxygen, little to no biodegradation of oxo-degradable plastics is expected in deeper landfill layers. In addition, there is insufficient evidence to indicate that oxo-degradable plastics will biodegrade in a reasonable timeframe in the marine environment (European Commission 2018).

Overall, there is a lack of significant evidence that biodegradable, compostable, biobased, and oxo-degradable plastics will fully degrade in natural environments (UNEP 2015; European Commission 2018, 2019). Further studies would assist in understanding the environmental impacts of different types of biodegradable, biobased, compostable and oxo-degradable plastic, particularly in comparison to conventional, petroleum-based plastics.

4.2 Fate in water

The proportion of plastics present in surface waters and sediments varies depending on the biological (e.g., attachment of bacteria/algae), physicochemical (e.g., plastic density), and hydrodynamic conditions (e.g., mixing of the water column) (Alimi et al. 2018). Factors such as wind, surface water circulation, temperature and salinity influence the distribution of microplastics (Zbyszewski et al. 2014; Corcoran et al. 2015; Anderson et al. 2016). 

In the aquatic environment, the rate of degradation of plastics is temperature-dependant, with degradation proceeding more slowly in cold water (Andrady 2015). Plastics found below the photic zone in the water column degrade very slowly, resulting in high persistence of plastics in the aphotic zone, particularly at the seafloor. In addition, biodegradation of plastics by microorganisms is negligible because of the slow kinetics of biodegradation at sea and the limited oxygen supply for these processes (Andrady 2015).

A study by Leonas and Gorden (1993) looked at disintegration rates of LDPE, PS, and a 2% ethylene-carbon monoxide polymer, as well as other blends in aqueous media. The results showed that while the ethylene-carbon monoxide polymer disintegrated^{Footnote 4} more rapidly than the other films evaluated, the aqueous environment significantly delayed, if not inhibited, the degradation of the other polymers.

Biber et al. (2019) studied the deterioration of different plastics in air and seawater. Macro-sized pieces of PE, PS, PET, and a commercial material marketed as degradable plastic were exposed to environmental conditions in air and water. All materials deteriorated more slowly in seawater than in air, likely due to reduced exposure to sunlight and thus reduced photooxidation in seawater. The authors found that PS showed the most rapid deterioration and is likely to break down into microplastics faster than the other materials evaluated, but that all materials tested did deteriorate to microplastics. Given the requirements for breakdown, it is expected that plastic items likely remain in seawater and that the formation of microplastics would occur in areas where plastic pollution is exposed to oxygen and UV radiation, such as intertidal habitats and at the water surface.

4.2.1 Sediment

Plastics may remain in benthic systems of lakes and rivers or be transferred along an altitudinal gradient towards marine ecosystems (e.g., oceans). As plastics move from source to sink, they interact with the physical, chemical, and biological environments in ways that depend on the characteristics of the plastics (e.g., density) (Windsor et al. 2019).

Besseling et al. (2019) found that microplastic concentrations on a volume basis are higher in sediments than in surface water. This can be explained by the settling of particles either as singular particles or in aggregated or fouled form. The authors also found that concentrations in beach sediments were higher than in subtidal sediments, which may be explained by the relatively low density of plastics compared to seawater, causing floating and suspended plastics to be washed ashore.

Sinking fecal matter from zooplankton that have ingested microplastics represents a mechanism by which floating plastics can be vertically transported away from surface waters and into deeper waters and the benthos, thus providing food for sediment-dwelling biota (Cole et al. 2016). Wieczorek et al. (2019) found that microplastics significantly altered the size, density, and sinking rates of zooplankton fecal pellets. In oceanic conditions, fecal pellets with reduced sinking velocities are more prone to consumption, fragmentation, and microbial degradation, resulting in their mineralization within the upper regions of the water column and therefore in reduced particulate organic matter export to deeper waters (Cole et al. 2016).

Fecal pellets containing microplastics that reside at the sea surface for a prolonged period are also more readily available for ingestion by other organisms, resulting in the trophic transfer of microplastics. Wieczorek et al. (2019) note that despite this, microplastics have been found in deep-sea sediments and benthic deep-sea organisms. Thus, an unknown proportion of microplastics are likely being transported to the seabed from fecal pellets where they become available to the benthos communities.

While sediment is largely expected to be a sink for macroplastics and microplastics (Eriksen et al. 2014; Woodall et al. 2014), there is significant mobilization and removal of microplastics in sediment during high flow events such as flooding (Hurley et al. 2018). Plastics in benthic sediments may be temporarily stored and remobilized by physical and biological processes. However, there is limited research on these mechanisms of plastic transport in aquatic systems (Windsor et al. 2019).

4.2.2 Impact of biofouling on aquatic distribution

Biofouling, also known as biofilm formation, is the coating of plastics with microorganisms, algae or plants. This process can lead to a loss of buoyancy and thus promote the sinking of microplastics to the bottom of the water body (Weinstein et al. 2016; SAPEA 2019). It has been hypothesized that phytoplankton aggregates act as potential sinks for microplastics (Long et al. 2015). Kaiser et al. (2017) found that the sinking velocities of PS particles increased by 16% in estuarine water and 81% in marine water after a six-week incubation period, which allowed for the particles to become coated with biofilms. The sinking of PE particles was not impacted by biofouling during 14 weeks of incubation in estuarine water, but in coastal water, their sinking velocity increased after six weeks. These results indicate that biofouling can enhance deposition of plastics to sediments and ocean beds (Kaiser et al. 2017). Further, Weinstein et al. (2016) indicated that biofilm formation on plastics decreases their UV transmittance, which could also inhibit the degradation of plastics in the environment.

4.3 Fate in soil

Although limited scientific information is available on the fate of plastics in the soil compartment, studies indicate that biodegradation can play a role in the fate of plastics in soil. Certain organisms, such as bacteria (Huerta Lwanga et al. 2018) or insect larvae (e.g., moths), can degrade plastics; however, this is not likely a relevant process in natural agroecosystems since they may not be naturally present in these environments (Ng et al. 2018). Alternatively, co-metabolism (i.e., the degradation of a compound in the presence of another compound used as a carbon source) is more likely to occur due to the abundance of carbon resources in soil (Ng et al. 2018).

The physicochemical state of plastics is also likely to be very dynamic in soil due to interactions with soil components, including organic matter (Ng et al. 2018). Interactions with certain pesticides can facilitate photodegradation or embrittlement of plastic particles (Schettini et al. 2014). As photo- and thermo-oxidative degradation pathways both rely on the combination of free radicals and oxygen, these processes only occur near or at the surface of soil (Ng et al. 2018). 

In a study by Cosgrove et al. (2007), PU was observed in soil at different organic carbon contents and different pH levels, and their fungal communities were compared. PU appeared to be highly susceptible to biodegradation in soil and was degraded almost completely after five months (Cosgrove et al. 2007; Eubeler et al. 2010). In another study, biodegradation in compost was investigated for irradiated ethylene propylene copolymers, LDPE, and isotactic PP films (Eubeler et al. 2010). The results showed that degradation increased with increasing irradiation time; however, after six months of exposure, LDPE was still the slowest sample to be degraded as measured by weight loss (Eubeler et al. 2010). Ohtake et al. (1995) found no evidence of biodegradation for PS, PVC and urea formaldehyde resin that had been buried under soil for over 32 years. Another study found that an LDPE bottle buried in shallow soil under aerobic conditions for over 30 years underwent degradation on the surface, but the inner part was almost unchanged (Ohtake et al. 1996).

Following the release of microplastics to the terrestrial environment, particles can be transported to surface water bodies by wind and water erosion or dispersed through ingestion by organisms (Maaß et al. 2017; Hurley and Nizzetto 2018). There is also the potential for microplastics to leach into groundwater aquifers due to downward drainage from soils (Re 2019). 

Soils are also expected to act as a major sink for plastic particles (Hurley and Nizzetto 2018). Microplastics are likely to be retained in soils for long periods of time due to factors such as vertical transport that draw the particles away from the surface, hindering degradation (Horton and Dixon 2017; Huerta Lwanga et al. 2017). Zubris and Richards (2005) studied fibres in soil as an indicator of the application of biosolids to land in the United States. The authors detected fibres in soil samples from field application sites up to 15 years after the application of sludge, and these data were corroborated with biosolids application records. Additionally, vertical transport of microplastics is possible via the movement of soil organisms (Huerta Lwanga et al. 2017; Maaß et al. 2017; Rillig et al. 2017) as well as agricultural processes (e.g., tilling), which can also cause physical damage to the particles (Ng et al. 2018).

4.4 Fate in air

Although research on the fate of microplastics in air is lacking, it is understood that air is likely an important pathway for the transport of microplastic particles (Dris et al. 2016). 

When released into the atmosphere, microplastics can become suspended, or further transported, due to their light weight (Horton and Dixon 2017). Suspension and dispersion of particles in the air is dependent on factors such as the size, shape and density of a particle (lighter, less dense polymers can be carried more easily), wind conditions, and precipitation, which can facilitate deposition on land or water (Dris et al. 2016; Prata 2018). 

Air currents and wind can transport particles long distances. Since air currents can be multidirectional, transport in air is less limited than transport in aquatic or terrestrial environments (Horton and Dixon 2017). For example, Allen et al. (2019) observed microplastic deposition in the French Pyrenees, a remote mountain catchment. Preliminary trajectory assessments showed that the microplastics had travelled up to 95 km from their source, indicating possible long-range transport. Microplastics have also been found in the Arctic Ocean in several studies. Lusher et al. (2015a) first reported the quantity of microplastics in surface and sub-surface Arctic polar waters. Subsequently, Bergmann et al. (2017) found large quantities of microplastics in Arctic deep-sea sediments, Kanhai et al. (2018) identified the abundance, distribution and composition of microplastics in sub-surface waters of the Arctic central basin, and Peeken et al. (2018) found microplastics in Arctic sea ice cores at five different locations and analyzed their content and composition. As with persistent organic pollutants, it is speculated that long-range atmospheric input of microplastics is one of several possible transport mechanisms (with others being prevailing currents and food-webs) to the remote Arctic Ocean (AMAP 2004; Kanhai et al. 2018; Peeken et al. 2018).

A number of studies have quantified microplastic presence in the atmosphere as well as in fallout (i.e., particles that settle on a surface during the sampling period). One study, for instance, observed atmospheric fallout of microplastics at a sampling site in a dense urban environment with a daily range of 2 to 355 particles/m² (Dris et al. 2016). A previous study by Dris et al. (2015) measured a total atmospheric fallout of 29 to 280 particles/m²/day in the urban Greater Paris region. In Dongguan City, China, the concentrations of microplastics in atmospheric fallout samples collected from three sites over a period of three months were 31, 33, and 43 particles/m²/day (Cai et al. 2017). Three different polymer types were identified in the microplastic samples (PE, PP, and PS), and fibres were the predominant shape of the microplastics sampled. In the metropolitan region of Hamburg, a median microplastic fallout concentration of 136.5 to 512.0 particles/m²/day was found across six sampling sites over a 12-week sample collection period (Klein and Fischer 2019). Of the microplastics detected, 95% were fragments, with fibres making up the remaining 5%. During periods of higher rainfall, Dris et al. (2016) observed a higher number of fibres in atmospheric fallout; however, there were likely other temporal and mechanistic factors at play, which the authors did not identify. 

In general, atmospheric concentrations of microplastics are likely to be correlated with population density, as human activities strongly influence the environmental release of microplastics. The fate and transport will depend on prevailing meteorological conditions, with long-range transport from urban source regions to remote locations highly probable.

5. Occurrence

This section reviews the available data on the occurrence of macroplastics and microplastics in aquatic and terrestrial environments and air, as well as in other matrices through which humans may be exposed to microplastics of environmental origin (namely food and drinking water). Occurrence in biota, with the exception of occurrence in food, is covered in Section 6. Where possible, Canadian occurrence data are presented. However, since Canadian occurrence data are often lacking, data from other areas around the world are also presented in many instances. 

The science assessment endeavours to discuss microplastic and macroplastic occurrence in the environment; however, peer-reviewed literature generally focusses more on microplastic rather than macroplastic occurrence.  Studies looking solely at macroplastic occurrence in the Canadian environment are often limited to data from litter cleanup initiatives as well as reports in the popular press.  Moreover, many studies on the occurrence of macroplastics in the environment are linked to effects such as entanglement or ingestion and much of this discussion is found in Section 6.

The absence of standardized methods and analytical techniques poses a significant challenge to quantifying microplastics in the environment. As a result, it is not possible to quantitatively characterize environmental or human exposure levels at this time. 

5.1 Environmental occurrence

5.1.1 Occurrence in the aquatic environment

Plastic pollution in the aquatic environment is summarized below with a focus on four compartments: shorelines (including the intertidal zone), surface waters, benthic zone (i.e., the bottom of a water body) and groundwater. 

As there are limited standardized procedures for quantifying microplastics in the aquatic environment, qualitative criteria were developed to identify studies that applied practices such as the use of controls, use of appropriate and clean glassware, and application of contamination avoidance measures. Further, the criteria used for studies, specifically on the occurrence of microplastics on shorelines and in surface water, selected studies in which microplastics were identified using an analytical method, such as Fourier transform infrared spectroscopy (FTIR), Raman spectroscopy, or pyrolysis gas chromatography mass spectrometry (GCMS). These qualitative criteria draw from the knowledge presented in Koelmans et al. (2019), which uses quantitative criteria for determining the quality of studies discussing the occurrence of microplastic in water samples. However, as acknowledged in Koelmans et al. (2019) and in Hermsen et al. (2018), these criteria are not an absolute judgement of the value of studies. Further, as illustrated in Gouin (2020) these metrics do not necessarily weigh all aspects of a study appropriately. For example, it is possible that studies that rely on visual identification of microplastics may score relatively higher when compared to studies that use analytical characterization. 

The science assessment reviews the current state of science regarding plastic pollution and acknowledges that uncertainties exist and high quality information is lacking in several study areas. As such, if any studies included in this report deviated from the above criteria, the limitation is explicitly mentioned in the text.

Shoreline

In an effort to remove litter from Canada’s shorelines, 21 300 cleanups have been organized by the Great Canadian Shoreline Cleanup (GCSC)^{Footnote 5}  across the country since 1994 (GCSC 2018a). Of the top 10 most common litter types collected during the 2018 Shoreline Cleanup, seven were either plastics or items containing plastic. Plastic items included cigarette butts, tiny plastics or foam, bottle caps, plastic bags, plastic bottles, straws, and food wrappers (GCSC 2018b). A total of 0.1 kt of litter was removed from Canadian shorelines in 2018 (GCSC 2018b). Figure 4 illustrates the contribution of plastics to shoreline litter collected during historical beach cleanup surveys of the Great Lakes.

It is worth noting that beach cleanups generally target larger intact and mostly intact debris, resulting in underestimations of smaller plastic debris (Zbyszewski et al. 2014). Moreover, brittle plastic materials may break into smaller fragments during analysis, and broken pieces may be quantified as microplastics rather than macroplastics, thus affecting counts (Esiukova 2017). Other methods of plastic quantification may underestimate the amount of microplastics, such as surveys on rocky shorelines. McWilliams et al. (2018) highlighted the need to further develop protocols and techniques to sample microplastics on rocky shores where marine litter may be caught between rocks and crevasses, thus increasing their lifetime on the shore. In addition, waves may grind macroplastics against rocky shores, accelerating their breakdown into microplastics. The authors conducted an accumulation survey of Fogo Island beach in Newfoundland and Labrador and found that 82% of marine litter collected from the rocky beach surface consisted of plastics, with 67% of litter being smaller than 1 cm³. 

McWilliams et al. (2018) also conducted a standing stock survey to assess the abundance of plastics at different depths of the beach on Fogo Island. This was done by first picking visible particles from the top layer of shoreline, followed by shovelling a 5 cm layer into a tray. Stratified sampling was performed to a depth of 20 cm. Across all layers, glass comprised 75.7% of litter, and plastics comprised 17.9%. More than 82% of the visually identified plastic particles sampled were macroplastics. Potential plastic particles were found throughout the different depths sampled, with the vast majority of smaller items found below the surface. Particles below the surface were found to be smaller and more abundant than particles on the surface. The surface accumulation survey in conjunction with the standing stock survey provides insight into how many particles would be missed by a survey that only considers particles on the surface.

In Lake Erie, Dean et al. (2018) collected 12 sediment samples from six beaches at the foreshore (between low- and high-water marks) and backshore (high-water mark to inland limit of beach). All samples contained microplastics, with a range of 50 to 146 particles/kg. Most of the backshore samples contained higher concentrations of microplastics than the foreshore samples. The dominant microplastic type was fibres, followed by fragments. It should be noted, that although precautions were taken against the contamination of samples from microplastics during processing, some samples were stored or sampled in PET jars or PVC liners, and sometimes new and unopened plastic containers were used in the field without prior rinsing (Dean et al. 2018). 

Proximity to industrial sources may be associated with higher concentrations of plastics (Zbyszewski and Corcoran 2011; Zbyszewski et al. 2014; Corcoran et al. 2015; Driedger et al. 2015; Ballent et al. 2016).  A review of existing shoreline studies by Driedger et al. (2015) indicated that higher concentrations of plastic debris in the Great Lakes region are generally found in areas with higher human and industrial activity. Zbyszewski et al. (2014) collected samples along the shorelines of Lake Erie, Lake Huron and Lake St. Clair. Macroplastics and microplastics were found along all shorelines. Additionally, Zbyszewski and Corcoran (2011) found that along the shoreline of Lake Huron, pellets represented the majority of the plastic collected. Further analysis found that most of the collected industrial pellets were PE and PP, similar to those produced by petrochemical companies (Zbyszewski et al. 2014).  The relative lack of plastic debris found on the north and west shores of Lake Huron in contrast to the southeast shore suggests that the pellets followed the cyclonic flow of surface water currents away from the region of Sarnia (Zbyszewski and Corcoran 2011; Zbyszewski et al. 2014). Similarly, Corcoran et al. (2015) collected 6172 plastic pieces from Humber Bay Park West beach on the northwest shoreline of Lake Ontario and found that industrial pellets were the most common type of plastic, followed by fragments. Excluding PS, which was only quantified by mass due to the large quantity collected, the plastics concentration was 21.8 items/m². The majority of pellets and fragments had accumulated within organic matter along the strandline. Several nearby tributaries pass through heavily industrialized areas before draining into Lake Ontario. The researchers observed similar types of plastic pellets in sampling sites along the tributaries and at the beach, suggesting a transport pathway (Corcoran et al. 2015). Ballent et al. (2016) found that the highest concentrations of microplastics in beach sediments along Lake Ontario were in the Greater Toronto Region, more specifically in an area made up of five watersheds that contained half of all plastic production facilities in the study region as well as 40% of the total population at the time. In this study, fragments were the predominant type of microplastic detected in beach sediment (average of 140 particles/kg dw), followed by fibres.

Munier and Bendell (2018) visually identified and collected plastic litter on the beach surface of Burrard Inlet in British Columbia. Of the 150 items collected, 144 were plastics, which were divided into seven major user groups: bags, car/bicycle parts, everyday items, food associated, packaging, functional use, and children’s toys. The majority of the plastics were items related to food consumption, such as cups, straws and forks, and packaging.

Single-use plastics are one of the most common types of macroplastics found on shorelines internationally. In Canada, 17% of collected shoreline litter consisted of plastic single-use food and beverage items (GCSC 2018b). Similarly, Earthwatch Europe (2018) found that single-use plastics are a significant category of plastic pollution in European freshwater environments. The 10 most prevalent macroplastics in European freshwater environments were plastic bottles (14% of identifiable plastic pollutants), food wrappers (12%), cigarette butts (9%), food takeout containers (6%), cotton bud sticks (5%), cups (4%), sanitary items (3%), smoking-related packaging (2%), plastic straws, stirrers and cutlery (1%), and plastic bags (1%) (Earthwatch Europe 2018). Cigarette butts rank high on both European and Canadian litter lists, with cigarette butts topping Canada’s 2018 Shoreline Cleanup for the highest abundance of litter and smoking-related litter making up 42.1% of the types of litter collected. In the northeast Atlantic, marine litter ranging from 2 to 30 cm was collected on beaches in the Azores, and plastic items accounted for 93% (26 321 items) of all litter. The collected litter consisted of 15.1% single-use items, 7.9% fishing-related items, and 71% fragments (Pieper et al. 2015). In the southern Caribbean, 42 585 litter items greater than 25 mm were collected at 10 locations on sandy beaches in Aruba. Of the litter collected, 89% (38 007 items) consisted of plastics. The collected litter was composed of 51% single-use plastics, of which 18% was bottle or container caps, 9% straws and 7% cigarettes. Additionally, 5% of all litter collected was fishing-related and 40% consisted of fragments of undiscernible origin (de Scisciolo et al. 2016). 

In Canada, fishing-related litter made up only 1% of the litter collected in the 2018 Canadian Shoreline Cleanup (GCSC 2018b). Additionally, plastic items related to fishing activity make up a significant amount of plastic pollution found on shorelines globally (Browne et al. 2010; Chen H et al. 2019; PAME 2019). Fishing-related litter is especially important in the Arctic, where most of the marine litter analyzed in the northern parts of Norway, the Barents Sea region, and the Arctic originated from fishing-related activities (Hallanger and Gabrielsen 2018). Fishing-related litter also accounted for 48% to 100% of the mass of litter on the beaches of Svalbard, Norway (PAME 2019). An average of 1 040 plastic items/km were collected in Iceland, corresponding to an average of 104 kg/km that mostly originated from Icelandic fisheries (Kienitz 2013). 

Plastics have been reported on shorelines around the world. Microplastics have been found on every Californian beach sampled by Horn et al. (2019), and fibres accounted for 95% of the microplastic items. Macroplastics have been found on beaches surveyed in Polynesia (Connors 2017) and on shorelines in East China (Chen H et al. 2019). In China, microplastics were collected on beaches adjacent to the Bohai and Yellow Seas, where flakes were the most abundant type of plastic (Zhou et al. 2018). On surveyed South African beaches, industrial pellets were the most abundant type of plastic (Ryan et al. 2018). Typically, pellets enter the environment via accidental spills on land or at sea, and weather conditions play a factor in industrial pellet accumulation, as well as the presence of beached organic materials (e.g., wood, weeds) in which they may become entrapped (Corcoran et al. 2015; Ryan et al. 2018). Microplastics have also been collected on beaches along the southeast coast of India (Karthik et al. 2018). Finally, both macroplastics and microplastics are widely found in the Arctic, despite its distance from industrialized and highly populated areas (PAME 2019). Refer to Section B-1 of Appendix B for further data on the occurrence of plastics on shorelines.

Surface water

Several studies have looked at the occurrence of plastic pollution in Canadian marine and freshwater bodies, with a focus on the Great Lakes. According to Driedger et al. (2015), areas with greater human and industrial activity in the Great Lakes region are generally associated with a higher abundance of plastic pollution in the adjacent Great Lakes basins.

In the Government of Canada’s 2015 Science Summary on Microbeads (ECCC 2015), several publications reporting on the presence of microplastics in marine and fresh waters were summarized. Building on this, a review of additional current literature on Canadian occurrence of microplastics in surface water is provided below.

Higher concentrations of plastics can be found near harbours or recreational areas and shipping routes (UNEP 2016). Hendrickson et al. (2018) studied microplastic occurrence in the surface waters of western Lake Superior. Sample sites were selected to include environments suspected to differ in microplastic distribution based on their proximity to presumed sources of microplastic pollution, such as WWTSs, urban shorelines and river outflows. On average, the estuary and harbour regions had the greatest abundance of microplastics, followed by open water sites and then nearshore sites. The average abundance for all sites was 37 000 particles/km². Fibres were the most abundant type of particle, followed by fragments, films, beads, and foams.

Anderson et al. (2017) found microplastics in all surface water samples collected from Lake Winnipeg at densities ranging from 66 788 to 293 161 particles/km². Microplastic densities in Lake Winnipeg were significantly higher than those reported for Lake Superior and Lake Huron, but were comparable to those of Lake Erie. There were no significant differences between nearshore and offshore sites. Fibres were the most common plastic type, whereas films and foams were the least common. In general, microfibres are one of the most common types of microplastic found in the aquatic environment (Anderson et al. 2017; Burns and Boxall 2018; Dean et al. 2018; Hendrickson et al. 2018; Collicutt et al. 2019; Corcoran et al. 2020). However, the distribution of microplastic type may also depend on the sample location as well as the method of quantification.

Grbić et al. (2020) quantified microplastics and other anthropogenic particles in Lake Ontario. Samples were taken from surface waters, wastewater effluent from three WWTSs, urban stormwater runoff, and agricultural runoff. PVC, PP and PE microplastics were found in all watershed types. PET and PE were the most predominant polymer types in surface water, PE in stormwater runoff, PET in wastewater effluent, and PVC and PP in agricultural runoff. 

Globally, plastics have been reported in fresh and marine surface waters, and extensive research has been done in marine surface waters. Macroplastics have been found in the Adriatic Sea, where plastic bags constitute nearly one-third of floating macroplastics (Zeri et al. 2018). Foamed PS items were the most frequently observed macroplastics surveyed in the South Pacific, North Pacific, South Atlantic, and Indian Oceans as well as around Australia (Eriksen et al. 2014). As plastic spreads throughout the world's oceans, it accumulates in subtropical gyres, such as the North Atlantic Subtropical Gyre and the North Pacific Subtropical Gyre, which is commonly referred to as the Great Pacific Garbage Patch (Eriksen et al. 2014; Poulain et al. 2019). Microplastics have also been found in Lake Michigan in the United States (Mason et al. 2016), in the Mediterranean Sea (de Haan et al. 2019) and in the Northwestern Pacific Ocean (Pan et al. 2019). Finally, macroplastics and microplastics have been found in Arctic surface water and in sea ice, and the majority of microplastics were fibres (Obbard et al. 2014; Lusher et al. 2015a; Peeken et al. 2018). Refer to Section B-2 of Appendix B for further data on plastic occurrence in surface waters.

Benthic zone

Microplastics have been reported in the benthic zone of Canadian waters. For example, in a study on the abundance and distribution of microplastics in surface sediments in Baynes Sound and Lambert Channel in British Columbia (Kazmiruk et al. 2018), particles visually identified as microplastics were found to be abundant in the sediment samples from all sampling locations. Microbeads were the most common type of microplastic found, with a maximum of 25 368 beads/kg sediment sampled at one site, followed by fibres and fragments. 

Ballent et al. (2016) quantified microplastics in Lake Ontario in nearshore, tributary, and beach sediment. Fragments and fibres were the dominant type of microplastic in the size range of less than 2 mm, and fragments and industrial pellets were the primary type of microplastic in the greater than 2 mm size range. Fibres were most abundant in nearshore samples, and pellets were present in all depositional environments, but not in sediment traps. PE was the most common type of polymer in the samples analyzed. The mean microplastic abundance was 760 particles/kg sediment. The highest abundances of microplastics were found in nearshore sediments, with 980 particles/kg dw, followed by tributary and beach sediments. Lake-bottom samples were also collected from Lake Ontario, with a total of 35 pieces being found in the two sample cores (Corcoran et al. 2015). No plastics were found in samples collected from depths greater than 8 cm.

Dean et al. (2018) examined microplastic occurrence in nearshore and tributary sediment along the shoreline of Lake Erie. Benthic sediment was sampled from Lake Erie nearshore locations, from the mouth of the Grand River, and from the Detroit River. Sediment samples were also collected from two northwestern Lake Erie tributaries and two northeastern tributaries. The concentration of microplastic particles in nearshore samples ranged from 0 to 391 particles/kg sediment, and fibres were the primary type of microplastic, followed by fragments. Tributary samples ranged from 10 to 462 particles/kg sediment with fragments dominating the samples, followed by fibres. A tributary sample from the Welland Canal, which is exposed to high shipping traffic and a sizable population, contained the largest concentration of microplastics. The passive sediment trap sample contained no microplastics, whereas the grab sample from the same location contained 390 particles/kg sediment (Dean et al. 2018).

Corcoran et al. (2020) investigated the distribution of microplastics in the Thames River in Ontario. Benthic sediment samples were collected from 34 locations along the river. All samples were found to contain microplastics, with an overall average concentration of 612 microplastics/kg dw sediment. Fibres were the most abundant (60%), followed by fragments (37%) and beads (3%). The most common polymer was PET for fibres and PE for fragments. Urban sections contained an average of 269 microplastics/kg dw, compared to rural sections which contained an average of 195 microplastics/kg dw. However, there was no significant influence of land use on the abundance of microplastics. Additionally, microplastics were found in the greatest abundance in samples with the finest sediment grain sizes and with the most organic debris.

Goodman et al. (2020) performed underwater video surveys of benthic debris in the Bay of Fundy. Surveys were conducted at 281 different locations, providing 33 hours of seafloor video footage. From the swept area, 47 items of debris were visually identified, 51% of which were categorized as plastic (71% of which were plastic bags). Abandoned, lost and discarded fishing gear, which included various plastics, comprised 28% of all debris. 

Globally, plastics have been reported in marine sediment, where they are typically dominated by microplastics. Macroplastics have been found in sediment in Argentina and the United Kingdom, and the dominant type of plastic was packaging and wrappers (Browne et al. 2010; Blettler et al. 2017). Macroplastics and microplastics have also been found in sediment in Italy, with fibres being the most abundant type of microplastics (Fastelli et al. 2016). Similarly, fibres were the predominant type of microplastic found in Croatia and in the Arctic (Sundet et al. 2016; Blašković et al. 2017; Renzi et al. 2019a). Microplastics found in river sediment in Shanghai consisted primarily of spheres, and the most dominant polymer was PP, similar to the situation in Hungary and on Rameswaram Coral Island, along the southeast coast of India (Peng et al. 2018; Vidyasakar et al. 2018; Bordós et al. 2019). Plastics have been collected from the Spanish Mediterranean seafloor (García-Rivera et al. 2018), the Arctic seafloor (PAME 2019), and the Pacific Ocean's Mariana Trench (Chiba et al. 2018). Refer to Section B-3 of Appendix B for further data on plastic occurrence in the benthic zone.

Groundwater

Groundwater is likely less vulnerable to microplastic pollution than surface waters (WHO 2019), although it has recently been hypothesized that microplastics from soils may be transported to and within aquifer systems (Re 2019). Currently, there is very little empirical data on the occurrence of microplastics in groundwater. 

Mintenig et al. (2019) investigated the presence of microplastics in drinking water derived from groundwater sources in the northwest region of Germany. Groundwater was supplied from wells at least 30 m in depth and microplastics over 20 µm were detected. Despite the use of very large volumes of water (1 000 L), very low microplastic concentrations were observed in groundwater, ranging from 0 to 0.007 particles/L, with a mean of 0.0007 particles/L. All identified microplastics were found to be small fragments between the sizes of 50 and 150 μm, with the predominant polymer types being polyester, PVC, PE, PA and epoxy resin.

A South African scoping study surveyed microplastics in groundwater from four boreholes in Potchefstroom (North West), South Africa (Bouwman et al. 2018). The mean microplastics concentration reported was 0.167 particles/L. From the microplastics identified, many of the fragments were in the lower size class range, below 600 μm. 

Panno et al. (2019) studied the occurrence of microplastics in springs and wells (<65 m) from two karst aquifers in the U.S. state of Illinois. Previous studies on the groundwater chemistry in these areas reported data suggesting input from septic effluent. The authors reported the presence of microplastics in 16 of the 17 water samples collected, with a median concentration of 6.4 particles/L and a maximum of 15.2 particles/L. Due to analytical limitations, the authors noted that it is possible that some particles they reported as plastic were actually fibres of natural origin.

5.1.2 Occurrence in the terrestrial environment

Litter

Information on the occurrence of plastic in litter is sparse; however, city litter audits have provided some data on the composition of litter in Canadian cities. For example, litter audits performed in Edmonton (2019), Toronto (2016) and Vancouver (2019) found that 32%, 31% and 46% of large litter, respectively, was composed of plastic (AET Group 2016; AET Group 2019; Dillon Consulting 2019). In these cities, large litter was defined as greater than 25.8 cm². Plastic items found in urban litter in these cities include cigarette butts, plastic films, straws and plastic packaging. Shoreline cleanups, such as the Great Canadian Shoreline Cleanup, also provide litter data, as summarised in Section 5.1.1.

Soil

The occurrence of plastics in soil is not as well studied as it is in water and sediment. Soil is an important environmental compartment in which to quantify microplastics as they may enter soils via plastic mulching or application of biosolids, among other sources. Given the lack of research on microplastic occurrence in soil, the criteria for selecting studies for this report were less stringent than for occurrence in surface waters.

Crossman et al. (2020) measured microplastics in biosolids from two suppliers and the soils from three agricultural fields in Ontario on which the biosolids were applied. One control field with no history of biosolid use was also sampled. Microplastics were found in all biosolid samples, with average concentrations ranging from 8 678 to 14 407 particles/kg dw. Overall, fragments were more predominant than fibres, comprising 63% to 73% of microplastics in biosolids. Average microplastic concentrations in soil ranged from 24 to 358 particles/kg. PE microplastics were the most abundant plastic polymer in both biosolids and soil. All fields that were previously treated with biosolids had higher soil pre-treatment microplastic concentrations compared to the control. The field with the greatest number of previous biosolid treatments had the highest pre-treatment soil microplastic concentration, suggesting the potential accumulation of microplastics from prior biosolid applications. Microplastic concentrations in soil increased significantly immediately after biosolid application in two fields, while the third showed a reduction. Only one field demonstrated a net gain in microplastics over the course of the study. Despite the high concentrations of microplastics that were applied to soil via biosolids, greater than 99% of those microplastics were not measured during soil sampling. Furthermore, there was an increase in the proportion of fragments in soil during and shortly after biosolid application, but the proportion of fragments eventually declined with a corresponding increase in proportion of fibres. This may indicate that fragments from biosolids are preferentially transported out of the soil matrix, but fibres are retained. 

In Germany, agricultural farmland was found to have 206 macroplastic pieces per hectare. The mean concentration of microplastics sized 1 to 5 mm was 0.34 particles/kg dry weight (dw) of soil (ranging from 0 to 1.25 particles/kg dw). The most common type of polymer for both macroplastics and microplastics was PE (67.9% and 62.50%, respectively) (Piehl et al. 2018).

Liu et al. (2018) found plastics in farmland soil around the suburbs of Shanghai, China. Macroplastic particles sized 5 mm to 2 cm were found at a concentration of 6.75 items/kg in shallow soil (0 to 3 cm) and 3.25 items/kg in deep soils (3 to 6 cm). Microplastic concentrations were 78.0 items/kg in shallow soil and 62.5 items/kg in deep soil. In general, Liu et al. (2018) found that topsoil contained higher concentrations of larger sizes of plastic particles. Fibres, fragments and films were the most common types of plastics and the majority of all plastics collected were PP and PE. Zhang and Liu (2018) also explored microplastic occurrence in arable land in southwestern China. The study area consisted of two cropped areas at the upstream and estuary of the Chai River, as well as a buffer zone, which was converted from cropland in 2009 to host indigenous trees. Plastic particles were found in all samples, ranging from 7 100 to 42 960 particles/kg, much higher than the concentrations measured by Liu et al. (2018). In addition, most microplastics were less than 1 mm in size and the dominant type of microplastic was fibres, constituting an average of 92% of samples. During wastewater treatment processes, microplastics can settle in sewage sludge, which can then be transferred to agricultural soils and used as fertilizer (Corradini et al. 2019). Corradini et al. (2019) sampled 30 agricultural fields in Chile with similar soil chemical and physical characteristics, but with different sludge application records over the past 10 years. The authors found high concentrations of microplastics in the soil and reported that microplastics accumulate in the soils with successive sludge applications. Scheurer and Bigalke (2018) found microplastics at concentrations up to 55.5 mg/kg (593 particles/kg) in floodplain soil samples in Switzerland, with a mean concentration of 5 mg/kg. Macroplastics sized 5 mm to 2.5 cm were also found but in much lower concentrations.

5.1.3 Occurrence in air

Indoor air

Limited data are available on exposure to microplastics in the indoor environment. Only two studies were identified in which indoor air was sampled, and three studies were identified in which fallout samples or settled dust samples (i.e., particles sampled from surfaces or vacuum cleaner bags) were collected to characterize microplastics in indoor air. Generally, particles were examined and counted microscopically and characterized by size, shape, and composition. However, collection and analysis techniques varied, and therefore comparison between studies is not possible.  

In the indoor environment, microplastics are more likely to occur in settled dust than in air, as they have a higher density than air (Henry et al. 2019). This route of exposure is particularly relevant to toddlers and young children, given behaviours such as crawling and hand-to-mouth activity. However, no data have been identified on partitioning of microplastics in indoor environments, and inhalation is therefore also considered a potential exposure route.  

Dris et al. (2017) looked at fibres in indoor air, indoor fallout, and settled dust in two apartments and an office in urban Paris. They found that approximately 33% of the fibres were synthetic, including PA, PP and PE. The authors reported air concentrations of 1 to 60 fibres per m³ (median 5.4 fibres/m³), and dust concentrations of 190 to 670 fibres per mg. The method was limited to fibres greater than 50 µm in length. However, there was an inverse relationship between the number of fibres and their size, suggesting that smaller fibres could be present in larger numbers. Fibre concentrations in indoor air were significantly higher than in outdoor air. 

Vianello et al. (2019) sampled indoor air in three apartments in Denmark and found that microplastics comprised 4% of the particles identified. The average number of microplastic particles in the samples was 9.3 per m³. Most (81%) of the microplastics were polyester; other polymers identified included PE, PP, and polyacrylonitrile (PAN). Both fibre and fragment shapes were observed, and the size limit for detection was 11 µm. Like Dris et al. (2017), Vianello et al. (2019) reported an inverse relationship between sample microplastic concentration and median of the size distribution.

Dust was examined from 39 homes in different locations across China (Liu C et al. 2019). PET was identified in all samples, and PC was found in 74% of samples. The method used could detect particles in the range of 50 to 2 000 µm; most microplastic particles were fibrous in shape. Synthetic polymers accounted for approximately 40% of the fibres collected, including polyester, PU, PA, PE, PP, and PAN. A concentration of 17 to 620 fibres per mg of dust was reported. The study authors also reported a concentration of PET in dust by mass (median of 27 µg per mg) and a concentration of PC in dust by mass (median of 0.005 µg per mg).

An earlier study (Schneider et al. 1996) looked at personal exposure to fibres at some European sites, using personal sampling pumps to collect airborne dust. The composition of fibres was not determined, but synthetic organic fibres may have included PE, PP, poly(vinyl alcohol), polyester, PA, and polytetrafluoroethylene (PTFE).

Cox et al. (2019) did a crude estimate of inhalation exposure to microplastics using the air concentrations of fibres from Dris et al. (2017) and Tunahan Kaya et al. (2018) (see outdoor air exposure section), assuming 33% of the fibres and particles were actually microplastics (Dris et al. 2017). Similarly, Prata (2018) used the data from Gasperi et al. (2015) to estimate the number of airborne microplastics that could enter the human lung each day. However, as discussed above, no quantitative estimate of exposure to microplastics from indoor air and dust was conducted for this assessment due to the limited number of studies available, the very small sample sizes, and the varying techniques and criteria applied for sample collection and particle characterization. 

Outdoor air

Only a few studies have investigated microplastics in outdoor air samples. The monitoring methods employ sampling techniques in which predetermined volumes of air are passed through filters onto which particles are collected. In addition, passive techniques that depend on atmospheric fallout onto a sampling surface or filter are used. Confirmation of microplastic particles among other particles is then completed using traditional methods. There are no Canadian data available, but limited studies were conducted in Europe, Asia and the Middle East. 

Dris et al. (2017) measured total fibre concentrations, including microplastic fibres, approximately 3 metres from the roof surface of an office building located roughly 10 km from the centre of Paris (four times throughout the year to account for seasonal variations). The concentration of fibres measured outdoors ranged between 0.3 and 1.5 fibres/m³ (median of 0.9 fibres/m³) and was significantly less than concentrations measured indoors within the office and at two residential sites in the same region. One sample collected on a rainy day in winter contained five times more fibres, suggesting that the rain caused fallout of the fibres. The methodology used in this study has a lower observation limit of 50 µm. However, the results revealed a more elevated number of particles in the smaller size fraction, suggesting that microplastics smaller than 50 µm could be present in greater numbers. 

The concentration of suspended atmospheric microplastics (SAMPs) measured in Shanghai ranged from 0 to 4.18 SAMPs/m³ (mean of 1.42 SAMPs/m³) (Liu K et al. 2019a). Microplastic fibres comprised 67% of the SAMPs, followed by fragments and granules (30% and 3%, respectively). The size and concentration of microplastics was shown to vary with altitude. The concentration of SAMPs was highest closer to the ground (1.7 metres), and lower at higher altitudes (33 and 80 metres). Larger sized particles (>5 000 µm) were also detected near ground level and not at higher altitudes. SAMPs were shown to represent 54% of the total particles collected and were comprised mostly of PET, PE, polyester and PAN. Poly(N-methyl acrylamide) (PAA) was the predominant SAMP at the highest altitude. It should be noted that rayon was included in the definition of SAMPs. Because this material is synthesized using cellulose, it is not always grouped with microplastics. This study estimates that the average adult in Shanghai inhales 21 microplastic particles per day.

The limited data on microplastics in outdoor air, measured in France and China, collectively identify an exposure level of approximately 1 microplastic particle per m³ of air. The primary exposure form is through microplastic fibres. However, there is significant uncertainty with regards to exposure to smaller microplastic particles, particularly those below 50 µm. In outdoor air, it is anticipated that most human inhalation exposures would occur near ground level and that concentrations would depend on many factors, including geographical proximity to outdoor microplastic sources, wind, temperature and precipitation (Prata 2018). Since people spend less time in outdoor or transit environments, they would be exposed to fewer microplastics outdoors than indoors.

A few studies have investigated the contribution of tire wear emissions to ambient levels of PM_2.5 and PM₁₀ (Panko et al. 2013, 2019; Kole et al. 2017; Kreider et al. 2019). In general, tire wear pollution data are sparse, available for a few locations, and estimates are indirectly calculated based on limited observations. Notwithstanding study limitations, a recent analysis by Panko et al. (2019) suggests that tire wear emissions contribute less than 1% to ambient levels of PM_2.5 and less than 3% to ambient levels of PM₁₀.

5.2 Occurrence in food and drinking water

5.2.1 Occurrence in food

Current knowledge of the occurrence of microplastics in food is limited. The point sources of confirmed microplastics in food are currently unknown, although microplastics likely enter the food chain through plastic waste breaking down in environmental matrices, such as water and air. For example, animal species consumed by humans may ingest microplastics from aquatic environments or become exposed via trophic transfer of microplastics from prey to predator (EFSA 2016; Toussaint et al. 2019). It is also possible for ambient microplastics in the air to settle on food items (Catarino et al. 2018; Li et al. 2018a). In a number of microplastics occurrence studies, contamination of laboratory control blanks and test samples by background and/or ambient air microplastics was reported as a methodological challenge (Mathalon and Hill 2014; Lachenmeier et al. 2015). 

Some research has suggested that food manufacturing, processing, and handling, as well as food packaging materials, may be potential point sources of microplastics in food (Karami et al. 2018; Oßmann et al. 2018; Schymanski et al. 2018). However, to date, there is no conclusive scientific evidence that food packaging materials, when used as intended (i.e., under normal conditions of use), are a source of microplastics in food or bottled water. Further studies are needed to determine whether food manufacturing, processing and/or handling, as well as food packaging materials, may contribute to microplastic concentrations in food. 

The majority of available data on findings of microplastics in foods pertain to analyses conducted internationally and, unless otherwise stated, are not Canadian specific data. Most studies have focused on investigating microplastic content in seafood, specifically fish and shellfish harvested from non-Canadian marine environments (EFSA 2016; FAO 2017; Lusher et al. 2017; Barboza et al. 2018; Toussaint et al. 2019). 

The available data for other animal species that may be consumed as part of the diet of Indigenous Peoples is summarized in Section 6. It is noted that the available research is limited to identifying macroplastics and microplastics from the perspective of animal health. It is not anticipated that consumption of gastrointestinal (GI) tracts, which would likely have the greatest amount of microplastics for marine mammals, fish and seabirds, is a major source of country food consumption. For example, according to the Inuit Health Survey, the most commonly consumed country foods include the flesh and organs of various mammals, birds and fish, but the survey does not indicate that the GI tract is usually consumed (Egeland and CINE 2010a,b,c). There is a need for research to quantify the presence of microplastics in the animal tissues and organs that are typically consumed. 

While some peer-reviewed studies report the presence of microplastics in certain foods, the particles were not confirmed as plastic, as the methodology employed relied on visual inspection or crude staining (Mathalon and Hill 2014; Desforges et al. 2015; Lachenmeiser et al. 2015; Liebezeit and Liebezeit 2013, 2014, 2015; Rochman et al. 2015; Wójcik-Fudalewska et al. 2016; Karlsson et al. 2017; Kosuth et al. 2018; Renzi et al. 2018). Given the lack of certainty that the particles reported in these studies are, in fact, microplastics, these results are not considered further in this report. 

The available data on the occurrence of microplastics in food, including bottled water, are summarized below, with further details available in Appendix C. 

Fish and shellfish

The presence of microplastics in the GI tract of over 150 fish species is well-documented, with microplastic content ranging in number from 0 to 20 microplastics per fish and ranging in size from 130 µm to 5 mm (Lusher et al. 2013; Campbell et al. 2017; EFSA 2016; FAO 2017; Barboza et al. 2018; Hantoro et al. 2019; Liboiron et al. 2018, 2019; Slootmaekers et al. 2019; Toussaint et al. 2019). There is significantly less information available on microplastic occurrence in fish muscle, which is the tissue of bony fish that is typically consumed (Karami et al. 2017a; Abbasi et al. 2018; Akhbarizadeh et al. 2018). The existing information indicates that microplastic concentrations in muscle tissue are lower than what has been reported in the GI tract of bony fish. The majority of whole fish samples (including fish muscle tissue and viscera) purchased from local markets in Malaysia did not contain any microplastics.  Of the small proportion of samples that did contain microplastics, concentrations ranged from 1 to 3 microplastics per fish (Karami et al. 2017a). Conversely, microplastics were detected in all analyzed fresh fish samples from the Persian Gulf, at concentrations ranging from 3.1 to 4.6 microplastics per fish (Abbasi et al. 2018) or 0.57 to 1.85 microplastics per gram of fish muscle tissue (Akhbarizadeh et al. 2018). Most microplastics were larger than 100 µm, with fragments and fibres being the predominant particle shapes in fish muscle tissue (Abbasi et al. 2018; Akhbarizadeh et al. 2018). 

 

Microplastics have been detected in a number of edible species of molluscs, including mussels, clams, oysters, scallops, and snails (Barboza et al. 2018; Toussaint et al. 2019). The most commonly investigated species of molluscs is the blue mussel, which was found to contain 0 to 10 microplastics per individual mussel or 0.2 to 2.9 microplastics per gram of meat (De Witte et al. 2014; Van Cauwenberghe and Janssen 2014; Li et al. 2015, 2018a; Van Cauwenberghe et al. 2015; Catarino et al. 2018; Toussaint et al. 2019). Similar concentrations of microplastics have been reported in clams, oysters, scallops, and snails (Van Cauwenberghe and Janssen 2014; Li et al. 2015; Naji et al. 2018; Su et al. 2018; Hantoro et al. 2019). Fibres and fragments were the most commonly detected shape, ranging in size from 5 µm to up to 4.7 mm (EFSA 2016; FAO 2017; Catarino et al. 2018; Li et al. 2018a; Naji et al. 2018; Su et al. 2018). The concentration of microplastics detected in mussels varies; with higher concentrations of microplastics observed in the tissue of mussels harvested from waters with higher environmental concentrations of microplastics (EFSA 2016; Li WC et al. 2016; FAO 2017; Hantoro et al. 2019). 

Occurrence data on the presence of microplastics in crustaceans is extremely limited. The average microplastic content in green tiger prawns sampled from the Persian Gulf was 7.8 microplastics per individual (muscle tissue and exoskeleton combined), with filamentous fragments measuring 100 to 250 µm identified as the most abundant type of microplastic (Abbasi et al. 2018). Conversely, microplastics were observed in the digestive tract, head, and gills of whole brown shrimp, but not in the abdominal muscle tissue of peeled brown shrimp, sampled from the Clyde Sea (Devriese et al. 2015). Microplastics have also been found in the guts of lobsters at concentrations of up to 0.80 mg per individual, with fibres being the most frequently observed shape (Murray and Cowie 2011; Welden and Cowie 2016). 

Other foods

The occurrence of microplastics has also been reported in a very small number of other foods, including honey, sugar, beer, and salt (EFSA 2016; Peixoto et al. 2019; Toussaint et al. 2019). One study reported that the majority of fibres in honey samples were naturally occurring cellulose fibres, with only a small portion of fibres confirmed to be PET by spectroscopy, but the number of PET fibres was not reported (Mühlschlegel et al. 2017). The remaining honey studies and all sugar and beer studies used a non-specific staining method to identify particles in the food items and thus, none of these particles could be confirmed as plastic (Liebezeit and Liebezeit 2013, 2014, 2015; Lachenmeier et al. 2015; Kosuth et al. 2018). 

 

A recent review of microplastics in salt reported that their presence in commercial salts was common, although microplastic concentrations varied considerably depending on the origin and type of salt (Peixoto et al. 2019). Sea salts contained the highest concentrations of microplastics, ranging from 0 to 19 800 microplastics per kg of salt (Yang et al. 2015; Iñiguez et al. 2017; Karami et al. 2017b; Gündoğdu 2018; Kim et al. 2018; Renzi and Blašković 2018; Seth and Shriwastav 2018). Concentrations in lake and rock/well salts were much lower, ranging from 0 to 800 microplastics per kg of salt and 0 to 204 microplastics per kg of salt, respectively (Yang et al. 2015; Iñiguez et al. 2017; Karami et al. 2017b; Gündoğdu 2018; Kim et al. 2018). In most studies of salt, microplastics less than 500 µm accounted for the largest proportion of detected microplastics, with fragments and fibres being the most abundant microplastic shape, regardless of salt type (Yang et al. 2015; Iñiguez et al. 2017; Karami et al. 2017b; Gündoğdu 2018; Kim et al. 2018; Renzi and Blašković 2018; Seth and Shriwastav 2018).

Bottled water

A few studies have evaluated the occurrence of microplastics in bottled water (Wiesheu et al. 2016; Kosuth et al. 2018; Mason et al. 2018; Oßmann et al. 2018; Schymanski et al. 2018; Zuccarello et al. 2019). In one study, microplastics were detected in 93% of bottled water samples purchased from 19 locations in nine countries outside of Canada, with an average concentration of 10.4 microplastics greater than or equal to 100 µm/L (Mason et al. 2018). The number of particles in the 6.5 to 100 µm size range were reported. However, spectroscopic analyses were not performed at this size range, and thus the particles could not be confirmed as plastic (Mason et al. 2018).

Microplastic concentrations are reported to vary across bottle type (i.e., plastic, glass or beverage carton) and intended use conditions (i.e., single-use versus multi-use bottles) (Oßmann et al. 2018; Schymanski et al. 2018). The highest concentrations of microplastics were reported in water from older multi-use plastic bottles, followed by glass bottles, newer multi-use plastic bottles, single-use plastic bottles, and beverage cartons (Oßmann et al. 2018; Schymanski et al. 2018). Approximately 78% to 98% of the microplastics detected in bottled water samples were between 1 and 5 µm, with less than 7% of microplastics greater than 10 µm (Oßmann et al. 2018). The point source of microplastics in bottled water is still unknown, and the variation in the reported microplastic concentrations does not seem to correlate with bottle type alone. This suggests that the origin of reported findings of some microplastics in bottled water may be environmental (i.e., from the source water and air as a result of secondary microplastics forming in the environment).

5.2.2 Occurrence in drinking water

A limited number of studies have measured microplastics in tap water, and even fewer are considered reliable due to concerns with quality assurance measures (WHO 2019). Average microplastic particle concentrations in tap water have been reported to range from 0.0007 to 628 particles/L (WHO 2019), and microplastics as small as 1 µm in size have been measured in drinking water (Pivokonsky et al. 2018). Due to the limitations of existing detection techniques, no information is available on the occurrence of particles below 1 µm in size. The most predominant polymer types detected were PET and PP in the form of fibres and fragments (WHO 2019). 

In a WHO-commissioned review, Koelmans et al. (2019) reviewed 50 studies on microplastics in tap water, bottled water and freshwater. The majority of the studies were missing at least one of nine critical aspects of quality assurance (Koelmans et al. 2019). Specifically, the authors noted uncertainties with the concentrations reported in many of these studies and concluded that any information presented on the presence of microplastics in water must be interpreted with this knowledge. Relevant studies on microplastics in tap water and freshwater are summarized below. See section 5.2.1 for a review of relevant bottled water studies.

Pivokonsky et al. (2018) examined raw surface water and treated drinking water for microplastics from three drinking water treatment plants (DWTPs) in urban areas of the Czech Republic. Drinking water samples were analyzed by scanning electron microscopy (SEM), FTIR and Raman spectroscopy. The results from this quantitative analysis indicated average concentrations of 338, 443 and 628 particles/L for drinking water at each of the respective DWTPs, with microplastics smaller than 10 μm accounting for up to 95% of particles retained. Although 12 different materials were identified, PET and PP were found to be the prevailing microplastics in treated water collected at two of the DWTPs, while PP and PE were most abundant in treated water collected at the third DWTP. Some of the limitations that may affect the overall quality and reliability of this dataset include the use of small sampling volumes and failure to take sufficient measures to control background contamination (i.e., wiping down surfaces and working under clean air conditions) (Koelmans et al. 2019). 

Strand et al. (2018) did not find significant concentrations of microplastics in tap water sampled from 17 different locations across Denmark sourced by groundwater. Samples were visually examined by stereomicroscopy for all particles greater than 100 µm displaying microplastic-like characteristics. Only a single sample concentration was reported above the level of detection (LoD) of 0.58 particles/L, at 0.6 particles/L. Chemical analysis by FTIR revealed that of the particles exhibiting microplastic properties, only 3% were confirmed to be microplastics, with the remainder identified as cellulose-like material (76%), as having poor spectra (10%), as having an unknown spectra (7%), or as protein-like material (4%). Polymer types were reported as PP, PS and PET. Given the very low level of each type of plastic polymer identified in the tap water samples, the authors caution against drawing conclusions on the origin of the plastic contamination. Additional tap water samples were collected to investigate the occurrence of smaller microplastics 10 to 100 µm, and chemical analyses were performed by FTIR. Only a single concentration of 0.8 particles/L was reported above the LoD of 0.3 particles/L, in the form of fragments comprised of PP, PET, acrylonitrile butadiene and PU. Despite the small sample volumes used in this study, the data presented was found to be among the most reliable studies on the occurrence of microplastics in drinking water (WHO 2019). 

In a study on tap water derived from the purification of groundwater in northwestern Germany, Mintenig et al. (2019) investigated the abundance of microplastics at different locations within the drinking water supply chain. Particles were characterized using FTIR imaging, and microplastics down to a size of 20 μm were identified. Results indicated a low level of microplastic contamination of tap water derived from groundwater, with concentrations in both raw and drinking water ranging from 0 to 7 x 10³ particles/L and a reported mean of 0.7 x 10^-3 particles/L. Microplastic particles identified were small fragments between 50 and 150 μm in size, with the predominant polymer types identified as polyester, PVC, PE, PA and epoxy resin. Although this study lacks some aspects of quality assurance, such as the use of clean air conditions and absence of positive controls (Koelmans et al. 2019), when assessed on key quality control criteria, it was found to score the highest of all tap water studies by the WHO (WHO 2019).

Two studies (Uhl et al. 2018; Kosuth et al. 2018) were identified but not considered reliable due to uncertainty about whether the methods used could accurately identify particles as plastic. In one study, no particles were observed in treated or distributed water in 24 DWTPs in Norway (Uhl et al. 2018). In another study, Kosuth et al. (2018) evaluated synthetic particles in tap water from 14 countries across five continents and found particles in 81% of samples, with the most abundant type being fibres. Concentrations ranged from 0 to 61 particles/L, with an overall mean of 5.45 particles/L.

Finally, the possibility exists that microplastic contamination could occur at some point in the water supply chain as a result of abrasion of water pipes containing plastic materials, or from membrane  filters made of polymers (Novotna et al. 2019). Further research is required to investigate this possibility.

5.2.3 Drinking water treatment

DWTPs provide a barrier against the introduction of waterborne microplastics in drinking water. The current literature, while limited, shows that drinking-water treatment can be effective at removing microplastics. However, given the lack of standardized methods for quantifying microplastics in water, further research is required in this area (Novotna et al. 2019). 

Drinking water treatment typically occurs via clarification or membrane processes. Clarification processes are the most commonly used methods for removing particles from drinking water and involve techniques such as coagulation, flocculation, flotation, and/or filtration (Novotna et al. 2019). Membrane processes involve the use of diffusion membranes (e.g., reverse osmosis) or porous membranes (e.g., microfiltration, ultrafiltration). Diffusion membranes allow only dissolved substances (such as ions and specific dissolved substances) to pass through, whereas porous membranes allow only particles of a certain size to pass (Crittenden et al. 2012). As most observed microplastics are above the membrane size thresholds for porous membranes (i.e., 0.1, 0.01 and 0.001 µm for micro-, ultra- and nano-filtration respectively), porous membranes have the potential to be very effective at removing microplastics (Crittenden et al. 2012). For example, a laboratory study by Ma et al. (2018) found complete rejection of PE microplastics by an ultrafiltration membrane. The type of drinking water treatment process may affect the efficiency of DWTPs in removing microplastics. However, further research is required to inform drinking water treatment optimization for microplastics. Pivokonsky et al. (2018) observed microplastic removal rates of between 70% and 82% for three DWTPs employing conventional coagulation, clarification, and filtration. In a study using groundwater, Mintenig et al. (2019) found no significant difference between source water and treated water, although microplastic concentrations were very low in both source and treated water, varying from 0 to 0.007 particles/L (Mintenig et al. 2019). 

Microplastic properties (e.g., size, shape, and surface properties), as well as water properties (e.g., pH and organic matter content) may also impact the efficiency of microplastic removal during different treatment processes. As microplastics are hydrophobic, adsorption of organic materials to the particles can occur, which can prevent their aggregation and thus make separation more difficult (Napper et al. 2015; Koelmans et al. 2016). Hydraulic forces can also break down large aggregates or particles themselves, creating smaller particles that may not be removed as easily during the clarification process (Jarvis et al. 2005). Ma et al. (2018) found that while pH and turbidity of the water had little effect on the microplastics removal efficiency, the microplastics themselves can actually influence the turbidity of water at sufficient concentrations. 

6. Impacts on environmental health

This section reviews data on the effects of both macroplastics and microplastics on environmental receptors. Each subsection begins with a discussion of occurrence in biota, followed by an overview of their effects.

6.1 Macroplastic

Plastic pollution can have various effects on organisms and their habitats, depending on the size and type of plastic, and the level of biological organization (Werner et al. 2016). In 2016, the Secretariat of the Convention on Biological Diversity (CBD) reported that a total of 817 marine species had been affected in some way by marine litter (CBD 2016), up 23% from the same assessment performed four years earlier. It also found that over 80% of this marine litter was plastic (CBD 2012, 2016). A literature review of 340 publications involving 693 species found that, globally, 92% of reported interactions between litter and species were related to plastic pollution (Gall and Thompson 2015). 

Rochman et al. (2016) conducted an extensive literature review of primary publications (283 papers) on marine litter (including macro- and micro-sized plastic pollution) published through to 2013. The authors compiled the perceived and demonstrated effects of litter and sorted them by levels of biological organization: suborganism, organism, population and assemblage. Micro-sized litter (defined as <1 mm in this study) accounted for 71% of the demonstrated impacts, while macro-sized litter (defined as >1 mm in this study) accounted for 29%. A further breakdown of these effects by level of biological organization shows that of the demonstrated impacts from macro-sized litter, the majority were classified as suborganismal, with the most common effects being seen in tissues (e.g., inflammation or lacerations) and organ systems (e.g., poor functioning). Of the demonstrated impacts at the suborganismal level, 78% were due to micro-sized litter, 74% of which were caused solely by plastics. Other demonstrated effects include effects on cells (e.g., necrosis, viability), in organs (e.g., change in size, lesions) and macromolecules (e.g., protein, DNA damage). All of these demonstrated impacts of macro-sized litter were found to be from plastic pollution. The remaining demonstrated effects were divided between the organismal level and the ecological level. At the organismal level, the main effect observed was death to an individual, whereas at the ecological level, the main effect was on assemblages (i.e., change in abundance or diversity of biota). The most common items reported to cause an effect were lost and abandoned fishing gear or other plastic items, such as rope, bags, straws and degraded fragments.

The adverse effects of macroplastic pollution include entanglement, ingestion, and impacts on habitat integrity (Gall and Thompson 2015; Rochman et al. 2016; Werner et al. 2016). 

6.1.1 Entanglement

Entanglement from macroplastics can occur from ropes, nets, cable ties, plastic bags, packaging bands and rings (such as for cans in bulk), and other string-like items (Werner et al. 2016). Observations of entanglement are reported more frequently than other impact pathways, likely due to its very visible nature (Werner et al. 2016). For example, Gall and Thompson (2015) found reported occurrences of entanglement for 30 896 individuals from 243 species. Of these reported cases, 79% were linked to direct harm or mortality, and the majority of these incidents involved plastic rope and netting. As well, Rochman et al. (2016) found that 29% of demonstrated impacts at the organismal level were caused by entanglement. The species most commonly impacted by entanglement events were marine invertebrates (75 species), seabirds (49 species), fish (27 species), and marine mammals (10 species). 

Entanglement in the marine environment is often due to “ghost fishing,” which occurs when lost, abandoned, or discarded fishing gear continues to catch fish in the ocean or on the seafloor (Hallanger and Gabrielsen 2018; PAME 2019). In the Arctic, old fishing-related products were found entangled with dead seabirds, dead and living Svalbard reindeers (Rangifer tarandus platyrhynchus), and seals (Hallanger and Gabrielsen 2018). In addition, Page et al. (2004) found the entanglement rates of Australian sea lions (Neophoca cinerea) and New Zealand fur seals (Arctocephalus forsteri) to be 1.3% and 0.9%, respectively, in 2002. These are some of the highest reported entanglement rates for all seal species. The authors estimated that 1478 seals die from entanglement events each year in Australia. Good et al. (2007) recovered 494 derelict fishing nets from Puget Sound and the Northwest Straits, along the coast of Washington, USA. Overall, 74% of the 7 539 organisms that were entangled in the derelict nets were dead, including marine invertebrates and vertebrates, of which 71% and 96% were recovered dead, respectively. All of the 123 birds and 16 mammals, including Harbor seals California sea lions and a Harbor porpoise, were recovered dead. In addition, a review of global data by Ryan (2018) reported that a total of 265 bird species were reported to be entangled in discarded plastics or other synthetic materials. Fishing gear was determined to be the cause of entanglement in 83% of species.

Votier et al. (2011) examined the use of macroplastics as nesting material by northern gannets (Morus bassanus) in Grassholm, Wales and assessed the associated entanglement events. Nests contained an average of 469.9 g dw of plastic and the preferred material used was synthetic rope. The authors estimate that, on average, 65.6 birds are entangled each year, with the majority being full-grown nestlings.

Large plastics such as bags, sheets, and films can also cover plants, sponges, and corals, affecting gas exchange and their photosynthetic capacities (Werner et al. 2016). This phenomenon, known as “smothering,” can lead to mortality of affected vegetation (Kühn et al. 2015). Rochman et al. (2016) found that 8% of deaths at the organismal level were due to smothering when examining demonstrated effects. Smothering by plastic pollution can also lead to sublethal effects in these organisms. To study the effects of smothering on cold-water corals (Lophelia pertusa), Chapron et al. (2018) used 10 x 10 cm pieces of LDPE to represent fragments of plastic bags, which have been seen covering polyps in the field. They observed a decrease in growth rates from 3.59 mm/year in control aquaria conditions to 2.51 mm/year in the test group exposed to macroplastics. The plastics may have acted as physical barriers to feeding, leading to impaired energy acquisition and slower growth rate. In addition, activity was 11% lower in coral exposed to macroplastics in comparison to control conditions after 7 days. However, activity was enhanced after 20 days, which the authors hypothesized to be a compensatory physiological response to enhance capture efficiency or a mechanism to cope with long-term low oxygen supply (Chapron et al. 2018). Macroplastic exposure also led to a noticeable decrease in feeding rates throughout the duration of the experiments.

Similarly, Qi et al. (2018) found that exposing soil to plastic films (1% w/w) had weak effects on the growth of wheat (Triticum aestivum). Plastic mulch films, comprised of 37.1% Pullulan (a polysaccharide), 44.6% PET and 18.3% polybutylene terephthalate (PBT), had stronger negative effects on wheat growth compared to the PE mulch. The authors note that this might be related to the presence of PET and PBT in the mulch, which have been shown in previous studies to have stronger negative effects on soil-plant systems than LDPE (Qi et al. 2018). However, exposure to both types of films inhibited wheat growth with respect to plant height at day 40 and shoot biomass at 2 months. The plants in both plastic mulch treatments also displayed fewer leaves, decreased leaf surface areas, and thinner stems. 

6.1.2 Ingestion

Ingestion of plastics is another pathway that can lead to potential adverse effects. Ingestion of plastic can be intentional (e.g., where an organism eats the plastic, mistaking it for food), or unintentional (e.g., where predators feed on prey that have ingested plastics). Filter-feeding or detritus-feeding species are especially prone to unintentional plastic ingestion (Werner et al. 2016). 

In the Mediterranean Sea, PE macroplastics were found in the gastrovascular cavity of 2 of 20 sampled jellyfish (Pelagia noctiluca) (Macali et al. 2018). Bernardini et al. (2018) also sampled 139 blue sharks (Prionace glauca) from the Mediterranean Sea. Blue sharks in the Mediterranean basin are categorized as a “Critically Endangered” species by the International Union for Conservation of Nature. Of the 95 adult blue sharks that were examined and had full stomachs, 24 contained plastic pollutants. Juveniles were also found to have a greater frequency of ingested plastics. In addition, macroplastics accounted for more than 70% of all plastic pieces. The majority of ingested plastic items were sheet-like (72.38%), followed by fragments (18.10%) and threadlike plastic items (5.71%), with the most common polymer found being PE. 

The GI tracts of Atlantic cod (Gadus morhua), European flounder (Platichthys flesus), common dab (Limanda limanda), Atlantic herring (Clupea harengus), and Atlantic mackerel (Scomber scombrus) caught from the North Sea and Baltic Sea were sampled for plastics by Rummel et al. (2016). Of the 290 investigated fish, 16 contained plastics (approximately 74% microplastics and 26% macroplastics). Macroplastics and microplastics were found in the GI tracts of 47.7% of the coastal fish and 2.4% of the offshore fish collected from Scottish marine waters by Murphy et al. (2017), or 29.7% (n=63) of all fish sampled. The mean number of plastic pieces found per fish was 1.8, with PA being the most common polymer. Choy and Drazen (2013) also found plastics in the stomachs of seven different species of pelagic fish from the central North Pacific Subtropical Gyre, many of which were macroplastics. 

Schuyler et al. (2014) conducted a global analysis of plastic ingestion in various sea turtle species and found that the most commonly ingested anthropogenic pollutants were plastics. Plot and Georges (2010) reported a field observation of an adult leatherback turtle that expulsed 2.6 kg of plastic pollutants, consisting primarily of plastic bags and plastic fragments. Plastics have also been found in green sea turtles (Chelonia mydas) (Özdilek et al. 2006; Stamper et al. 2009). 

Lusher et al. (2015b) studied two adult and one juvenile True’s beaked whales (Mesoplodon mirus) that were found stranded on the coast of Ireland. Analysis of the contents of their stomachs and intestines revealed that both adults appeared to have ingested macroplastics, but it could not be determined whether the whales died as a direct consequence of plastic ingestion (Lusher et al. 2015b). Marine litter was also found in the stomachs and intestines of 26 out of the 175 (approximately 15%) dead Magellanic penguins (Spheniscus magellanicus) collected from the Brazilian coastal zone, roughly 58% of which was plastics (Brandão et al. 2011). 

Gall and Thompson (2015) reported occurrences of marine litter ingestion for 13 110 individuals of 208 species, and Kühn et al. (2015) reported that the number of species known to ingest plastics increased by approximately 87% from 1997 to 2015 (177 to 331 species) and that marine litter ingestion has been recorded in 50.4% of marine mammal species, 40.4% of seabird species, and 100% of turtle species. However, cases of plastic ingestion leading to direct harm or death is less frequent in comparison to entanglement. Gall and Thompson (2015) found that only 4% of reported cases of ingestion resulted in direct harm or death. In contrast, Rochman et al. (2016) found that 63% of deaths were due to ingestion of marine litter. Specifically, demonstrated impacts from ingestion were observed in marine mammals (two species), sea turtles (one species), seabirds (one species), and marine invertebrates (two species). 

Ingestion of plastics by organisms has been shown to have consequences from several pathways. Current literature shows that the most clear adverse effects from plastic ingestion is the blockage of intestinal systems, preventing feeding and thus leading to possible starvation. For example, a common dolphinfish (Coryphaena hippurus) caught in the Western Equatorial Atlantic had a large plastic bowl measuring 99.57 cm² in its stomach (Menezes et al. 2019). Researchers suggested that the bowl was likely blocking its digestive tract, leading to starvation. A study by Pierce et al. (2004) reported plastic ingestion by a male northern gannet (Morus bassanus) and a female greater shearwater (Puffinus gravis). Both birds had blockages of the pylorus, which prevented feeding, leading to starvation and death. Ulcerations near the pylorus were also seen in the northern gannet, which matched up exactly with the shape of the bottle cap found in its esophagus that was thought to have been dislodged from the gizzard.

Ingested plastics can also damage organs and intestinal systems. Brandão et al. (2011) observed a dead Magellanic penguin (Spheniscus magellanicus) whose stomach had been perforated by a plastic straw. Jacobsen et al. (2010) studied two sperm whales (Physeter macrocephalus) post-mortem, both of which had netting, fishing line, and plastic pollutants such as bags in their stomachs. The cause of death in both whales was suspected to be gastric impaction, as one whale had a ruptured stomach and the other was emaciated. Stamper et al. (2009) observed an emaciated green sea turtle (Chelonia mydas) floating off the coast of a Florida beach. The turtle displayed signs of cachexia, lethargy, increased buoyancy, obstipation, and anorexia. Radiographs confirmed the presence of plastics in the GI tract, hindering regular function. After the removal of 74 foreign objects (including latex balloons, string, nylon rope, and soft and hard plastics) via enemas, the turtle showed improvements in its health, appetite, and behaviour. The authors note that this demonstrates a cause-and-effect relationship between plastic ingestion and morbidity in organisms (Stamper et al. 2009). 

6.1.3 Habitat integrity and rafting (organism transport)

The presence of plastic pollution in water bodies can also pose potential problems for ecosystem function, biodiversity, and habitat integrity (Werner et al. 2016). An increasing amount of plastic pollution in surface waters has the potential to act as a stressor on ecosystem dynamics and habitat integrity (CBD 2012). 

Plastics can be effective transport mediums due to their potential for surface adhesion and to the low density of certain types of plastic and can potentially accentuate transport of organisms or other organic matter, a phenomenon known as “rafting” (Werner et al. 2016). This process can also occur with naturally occurring material such as wood, but the increasing prevalence of plastic pollution in surface waters increases the likelihood for organisms to be transported, which can pose a threat to the receiving environment. Gall and Thompson (2015) identified 34 reports of organisms rafting on marine litter, including packaging, fragments, and intact items (plastic or otherwise). Of the 259 total species described in these reports, six were listed as being invasive (i.e., non-native). However, the authors note that this is likely an underrepresentation (CBD 2012; Gall and Thompson 2015). The transport of non-native species is a particular concern, as they have the potential to negatively impact the structure of other well-established ecosystems by becoming predators to native species and/or outcompeting them for resources, leading to a loss of biodiversity (Werner et al. 2016). Non-native species could also transport diseases to which native species have not previously been exposed and could alter the genetic diversity in the ecosystem. Furthermore, plastic pollutants can also act as an artificial habitat for the colonization and growth of microorganisms that can affect species assemblage (Werner et al. 2016).

Katsanevakis et al. (2007) studied the impacts of marine litter on the abundance and community structure of epibenthic megafauna in the Aegean Sea. They demonstrated that an increase in marine litter caused a marked and gradual increase in both the total abundance and number of species, changing the structure of the megafaunal community. This was attributed to the fact that the litter was able to provide refuge for mobile species and to act as a colonization site for hard-substratum sessile species. This change in dynamics can have significant long-term effects on the ecosystem, such as altered predator-prey dynamics. 

6.2 Microplastic

There are no standardized methods for testing the effects of microplastics. Currently, concentrations of microplastics used in effect studies are much higher than those measured in the environment (Burns and Boxall 2018). Furthermore, effects studies focus on particle sizes much smaller than those currently sampled for in the environment (SAPEA 2019). Particle concentration can also influence toxicity, as higher concentrations are expected to overwhelm biological clearance mechanisms and cause responses that are not otherwise observed at lower doses (WHO 2019). Results from Pikuda et al. (2019) indicate that preservatives in commercial plastic formulations, rather than the plastic particle itself, may be responsible for the observed acute toxicity to test organisms. However, the washing of test particles is not currently standard practice and therefore this was not considered in the above criteria.

For the purposes of this report, the following criteria were used to select the studies: the study reported details of the analytical techniques, the study reported the type of plastic used (i.e., polymer type, size, shape, virgin vs. aged), and the study monitored and reported measured concentrations that were similar to the nominal (i.e., theoretical) concentrations. Similar to the environmental occurrence section, these qualitative criteria draws upon the quantitative criteria presented by Hermsen et al. (2018) for the determination of study quality in papers examining the ingestion of microplastics by biota. However, as acknowledged in this paper and in Koelmans et al. (2019), these criteria are not an absolute judgment of the value of studies, because not all aspects of studies could be captured in the scoring systems. As such, if any studies included in this report deviated from the above criteria, the limitation is explicitly mentioned in the text. Furthermore, studies in this section were selected in order to cover a variety of organism types and effects.

6.2.1 Uptake, ingestion, and egestion

Microplastics have been found in many species, including invertebrates, fish, turtles, mammals, and birds. Given the lack of standardized methods for quantifying occurrence in biota as well as the limited data on occurrence in Canadian species, criteria for selecting reliable studies (e.g., studies that used an analytical method to identify microplastics) were identified but many studies did not meet these standards. Moving forward, it is recommended that a standardized method for quantifying microplastics in biota be developed.

A review by Provencher et al. (2017) showed that the literature on global macroplastic and microplastic ingestion in marine vertebrates is dominated by seabirds and that there is an increasing number of reports in fish, turtles and mammals each year. Fibres and fragments are the most common microplastic types found in organisms (Burns and Boxall 2018). For example, Beer et al. (2018) visually identified microplastics in 20% of the 814 fish they studied in the Baltic Sea, with 93% of these being fibres. Collicutt et al. (2019) determined by light microscopy that over 90% of the microplastics they found in juvenile Chinook salmon (Oncorhynchus tsawytscha) were fibres. 

As in the case of macroplastics, several factors can affect the intake and ingestion of microplastics by organisms. In laboratory studies, Scherer et al. (2017) demonstrated that co-exposure of microplastics with algae significantly decreased ingestion of microplastics by Daphnia magna. Weber et al. (2018) demonstrated that exposure concentration and age of the freshwater amphipod Gammarus pulex affected its microplastic body burden. Feeding selectivity of biota is also thought to be a driving factor for microplastic ingestion: non-selective filter feeders are more prone to direct microplastic uptake, whereas more specialized feeders will uptake microplastics indirectly through ingestion by their prey (Scherer et al. 2018). Uptake of microplastics via prey ingestion is discussed further below. Select reported ingestion events are outlined below, with Canadian and global examples. 

Liboiron et al. (2019) studied the GI tracts of three fish species commonly used for human consumption on the island of Newfoundland: Atlantic cod (Gadus morhua), Atlantic salmon (Salmo salar) and capelin (Mallotus villosus). The frequency of occurrence of macroplastic and microplastic ingestion by Atlantic salmon and capelin was 0% for specimens collected between 2015 and 2016 (a total of 419 fish). In Atlantic cod examined during the same period, the frequency of occurrence of plastic ingestion was 1.68%. These results are consistent with a previous study by Liboiron et al. (2018), in which 134 silver hake (Merluccius bilinearis) from the south coast of Newfoundland were studied and found to have a 0% frequency of occurrence of plastic ingestion. 

In a study of microplastics in the GI tract of juvenile Chinook salmon on the east coast of Vancouver Island, Collicutt et al. (2019) found that 59% of the specimens examined contained at least one plastic particle, with an average of 1.15 microplastic pieces per individual. It should be noted that plastic identification was not confirmed using an analytical method other than visual identification using light microscopy. 

In a study of microplastics in fish from a prairie creek downstream from a WWTS in Regina, Saskatchewan (Campbell et al. 2017), five species of fish were collected: fathead minnow (Pimephales promelas), northern pike (Esox lucius), white sucker (Catostomus commersoni), emerald shiner (Notropis atherinoides), and five-spine stickleback (Eucalia inconstans). Of the 181 fish sampled, 73.5% had between 1 and 20 microplastics in their GI tracts. The number of microplastics varied significantly between the five species sampled. This inter-species variation is hypothesized by the authors to be attributable to differences in feeding habits. The northern pike, an apex predator, had the highest proportion of sampled fish with microplastics present in their GI tracts at 83.3%, while the fathead minnow had the lowest, at 50.0%. The authors acknowledge that characterization of plastics using spectroscopic identification methods was not performed in this study. However, a hot needle was used to test whether the suspected plastic particles melted, to confirm that the particle was plastic (Campbell et al. 2017). It should be noted that some types of plastic will not melt under these conditions (i.e., thermosets).

O’Hara et al. (2019) conducted a study on the seasonal variability of exposure of Cassin’s auklets (Ptychoramphus aleuticus) to microplastic pollution. Following a series of storm events, 707 carcasses were found on the beaches of Vancouver Island and Haida Gwaii in British Columbia. A total of 85 carcasses were collected for examination, and plastics were found in the stomachs of 40% of the birds. Macroplastic and microplastic pieces in the stomachs of the birds were visually identified and separated. The average number of plastic pieces ingested per bird was 1.6, with an average mass of 0.0085 g, and one outlier ingested 61 pieces of plastic. Furthermore, ingested plastics were predominantly microplastics (86.6%). There was no significant difference between the number of pieces ingested by age, sex, or health condition of the bird (O’Hara et al. 2019). Similarly, Poon et al. (2017) studied plastic ingestion by Northern fulmars (Fulmarus glacialis) in the Canadian high Arctic. None of the stomachs of Northern fulmars sampled in 2013 contained more than 0.1 g of visually identified plastics. Provencher et al. (2018a) demonstrated that Northern fulmars excreted microplastics via their guano and found that the number of pieces of plastic in the gut was positively related to the number of microplastics in the guano. 

Plastics have been identified in organisms from several regions of the world. Representative studies are presented below to demonstrate that microplastic ingestion by biota occurs globally. An exhaustive review has been conducted by Provencher et al. (2018b).

Microplastics have been found in gudgeons (Gobio gobio) in Flemish rivers in Belgium (Slootmaekers et al. 2019). Gudgeons from 15 rivers at 17 locations were sampled to study the occurrence of microplastics in their intestines. Microplastic contamination was found in four of the rivers studied. Of the 78 fish examined, 9% contained microplastic particles in their intestines, and only one fish had ingested more than one particle. A total of 16 suspected plastic particles were extracted from all sampled fish; however, only eight particles were identified to be plastic following μ-FTIR analysis. Overall, seven different polymers were identified: ethylene-vinyl acetate copolymer, PP, PET, PVC, cellophane, polyvinyl acetate and PA (Slootmaekers et al. 2019).

In the heavily industrialized city of Tuticorin, India, Kumar et al. (2018) investigated the occurrence of microplastics in Indian mackerel (Rastrilliger kanagurta) and honeycomb grouper (Epinephalus merra) on the southeastern coast. Of the 40 fish sampled, 12 had plastic particles in their intestines. FTIR analysis revealed that the particles were PE and PP. Fibres constituted 80% of particles, whereas fragments constituted the remaining 20%. 

While the ingestion of microplastics has been widely demonstrated, egestion has also been shown to be possible in some organisms. For example, Grigorakis et al. (2017) found that goldfish (Carassius auratus) have efficient gut clearance of microbeads and microfibres: the time required for 90% clearance was 33.4 hours. Mazurais et al. (2015) found complete egestion of PE microbeads from European seabass (Dicentrarchus labrax) larvae after 48 hours. In invertebrates, significant microplastic egestion was seen in studies by Chua et al. (2014), Blarer and Burkhardt-Holm (2016), Frydkjær et al. (2017), and Hämer et al. (2014). In Hyalella azteca, an amphipod crustacean, microplastic fibres were found to be more slowly egested than microbeads during acute exposure; however, both were able to be completely egested (Au et al. 2015). 

6.2.2 Ecotoxicological effects

Despite the ability of some organisms to egest plastic particles, microplastics have been shown in the current literature to have adverse effects on organisms. In their respective literature reviews, Rochman et al. (2016) and the Joint Group of Experts on the Scientific Aspects of Marine Environmental Protection (GESAMP) (2016) reported that, in the laboratory, the effects from micro-sized litter (consisting primarily of plastic) were overwhelmingly seen at the suborganismal level. The predominant observed effects at this level were in macromolecules, cells, and tissues and can include inflammation and changes in gene expression (Rochman et al. 2016; GESAMP 2016). The remaining demonstrated effects were at the organismal level, primarily due to individual deaths (Rochman et al. 2016). In addition, Foley et al. (2018) conducted a meta-analysis of 43 papers published before October 2016 and observed that while effects from microplastic exposure were highly variable across taxa, the most consistently reported effect across both marine and freshwater taxa was a reduction in the consumption of natural prey. 

In the Government of Canada’s 2015 science summary on microbeads (ECCC 2015), 130 publications on the fate and effects of microplastics were examined and reviewed. Several key studies were summarized in the assessment report. The report noted a scarcity of information on long-term and multigenerational effects of microbeads; however, short-term and direct effects are well described. Physical effects were identified as the primary driver for effects to organisms. Some examples of effects in organisms from microbead exposure that have been described in literature include: decreased survival and fecundity (Lee et al. 2013), decreased reproduction from impedance of feeding behaviour (Cole et al. 2015), liver stress (Rochman et al. 2013), altered gene expression (Rochman et al. 2014), and possible genotoxicity in the form of DNA damage (Avio et al. 2015). Au et al. (2015) found that acute exposure to microfibres produced greater toxicity (due to physical effects) to Hyalella azteca than spherical beads, with 10-day median lethal concentration (LC₅₀) values of 71.43 microfibres/mL and 4.64 x 10⁴ microbeads/mL, respectively. Hämer et al. (2014) observed no impact on survival, growth, and intermolt duration in isopods (Idotea emarginata) following chronic exposure to microplastic particles of multiple forms. More detailed summaries of these studies can be found in ECCC (2015). 

In its proposal for a restriction on intentionally added microplastics, the European Chemicals Agency (ECHA) reviewed and summarized 25 influential scientific papers on the ecotoxicological effects of microplastics (ECHA 2019). The papers include data that overlap with those from studies cited in ECCC (2015). Experimental data cited by the ECHA in its proposal that were not discussed in ECCC (2015) are summarized briefly below. For more detailed summaries of these studies, please refer to ECHA (2019).

• Earthworms (Lumbricus terrestris) showed a decrease in growth rate with exposure to high concentrations of LDPE particles (<150 μm; 28, 45, 60% dw), but reproduction was not affected (Huerta Lwanga et al. 2016).
• Zebrafish (Danio rerio) exposed to PS microspheres (5 μm) exhibited inflammation, lipid accumulation in liver, oxidative stress, and altered metabolomics profiles (Lu et al. 2016).
• European sea bass (Dicentrarchus labrax) feeding on PVC pellets (<0.3 mm; 1.4% bw) had significant structural damage to the intestine (Pedà et al. 2016).
• Daphnia magna that ingested PE particles (1 μm; 12.5 to 200 mg/L) experienced immobilization that increased with concentration and time following 96-hour exposure (Rehse et al. 2016).
• Pacific oysters (Crassostrea gigas) exposed to PS spheres (2 and 6 μm; 0.023 mg/L) had significant reductions in oocyte number, oocyte diameter, sperm velocity, and larval development of offspring following two-month exposure (Sussarellu et al. 2016).
• Mussels (Mytilus edulis) and lugworms (Arenicola marina) exposed to PS microspheres (10 μm, 30 μm, 90 μm; 110 particles/mL seawater for mussels, 110 particles/g for lugworms) showed increased metabolism, but no adverse effects on energy allocation (Van Cauwenberghe et al. 2015).
• Common shore crabs (Carcinus maenas) feeding on PP microfibres (1 to 5 mm in length; 1% plastic) showed a decrease in food consumption rates over time and a drastic reduction in energy available for growth, with minimal lasting consequences (Watts et al. 2015).
• Marine worms (Arenicola marina) had reduced feeding activity and reduced available energy reserves from ingestion of unplasticized PVC treatments (130 μm mean diameter; 5% by weight) (Wright et al. 2013). 

A review of additional current literature on ecotoxicological effects of microplastics is provided below for each environmental compartment of interest. Relevant studies are outlined in the sections below, with more detailed summaries provided in Appendix D, including information on the size, concentration and polymer type of the particles. Due to physicochemical similarities, information on primary microplastics was used as surrogate information where information on secondary microplastics was not readily available.

Water

The aquatic environment and marine organisms in particular, have been the focus of much of the ecotoxicological research on plastics (SAPEA 2019). In freshwater studies, invertebrates have been the focus of research on sensitivity to microplastic exposure (Adam et al. 2019). 

Studies on the effects of plastics on organisms in both freshwater and marine environments are presented below, by level of biological organization. 

Vertebrates

Yin et al. (2018) exposed the fish species Sebastes schlegelii to PS spheres and observed a reduction in foraging time and swimming speed, an increase in shoaling behaviour, and a feeding time of almost twice that of the control. Goldfish (Carassius auratus) exposed to ethylene vinyl acetate fibres, PS fragments, and polyethylene acrylate pellets also exhibited sublethal effects such as weight loss, histological changes to the GI tract and intestines, inflammation of the liver, and physical damage to the jaw, including incisions from chewing fragmented particles, but no mortality (Jabeen et al. 2018). 

Similarly, in zebrafish (Danio rerio) exposed to PA, PE, PP, and PVC microplastics, no significant difference in lethality was observed; however, microplastics caused intestinal damage such as cracking of villi and splitting of enterocytes (Lei et al. 2018a). A study by Qiao et al. (2019a) reported similar findings when using PS microplastics, such that zebrafish exposed to virgin PS beads had significant intestinal damage, inflammation, oxidative stress, and altered gut microbiomes. 

At the molecular level, Qiang and Cheng (2019) found that exposure to PS microplastics induced upregulated expression of inflammation and oxidative stress-regulated genes in zebrafish larvae. S. schlegelii showed a significant reduction in crude protein and lipid contents and had black bile in their gallbladders, indicating GI function disorder resulting from accumulation of PS spheres in their intestinal tract (Yin et al. 2018). 

Conversely, several current studies report no significant effects on vertebrates for any of the endpoints measured. De Felice et al. (2018) exposed tadpoles of African clawed frog (Xenopus laevis) to PS microplastics and found no significant effects on mortality, body growth, or swimming activity during their early life stages, despite observing microplastics in the digestive tracts of all exposed tadpoles. Further, Ašmonaitė et al. (2018) observed no significant histological effects or inflammatory responses in rainbow trout (Oncorhynchus mykiss) exposed to PS microplastics, and Jacob et al. (2019) observed no effects on foraging or predation avoidance in coral-reef fish (Acanthurus triostegus) exposed to PS microbeads. Dietary exposure to PVC, PA, PE, and PS microplastics also did not affect stress responses, growth rate, or induce pathology changes in seabream (Sparus aurata) (Jovanović et al. 2018).

Invertebrates

In cladocerans (Daphnia spp.), increased microplastic concentrations led to increased mortality (Aljaibachi and Callaghan 2018; Martins and Guilhermino 2018; Pacheco et al. 2018). However, Jaikumar et al. (2018) suggest that mortality might also be temperature-dependent. Martins and Guilhermino (2018) further observed that microplastic exposure could have transgenerational effects in D. magna. Females descending from groups exposed to microspheres showed reduced growth, reproduction, and population growth rates up to the F₃ generation, indicating that complete recovery from chronic exposure may take several generations for this species. Tang et al. (2019) found no mortality in Daphnia exposed to PS particles, but observed a reduction in body growth rate and increased transcription of arginine kinase and permease (enzymes involved in oxidative defence and energy production).

Freshwater crabs (Eriocheir sinensis) exposed to PS microspheres similarly showed a decrease in weight gain, reduced hepatosomatic index, and several biochemical effects, such as an increase in transcription of genes involved in the oxidative stress response and anti-inflammation pathways (Yu et al. 2018). 

Similarly, Jeong et al. (2017) found that exposing the marine copepod Paracyclopina nana to PS microbeads increased antioxidant enzyme activity in a size-dependent manner. A previous study by Jeong et al. (2016) found similar results when studying the monogonont rotifer Brachionus koreanus: several antioxidant enzymes showed increased activity in rotifers exposed to PS microbeads, indicating a defence mechanism against oxidative stress.

Beiras et al. (2018) studied rotifers as well as the crustacean Tigriopus fulvus and determined lowest observed effect concentrations (LOECs) of 0.01 mg/L for rotifer immobility and 1.0 mg/L for rotifer and crustacean mortality, using PE particles. 

For the freshwater invertebrate Gammarus pulex, Weber et al. (2018) found no significant effects on juvenile survival, development (molting), metabolism, or feeding activity following chronic exposure to PET. Another study by Redondo-Hasselerharm et al. (2018) exposed G. pulex to PS microplastics. While the survival of G. pulex was not affected, the organisms experienced a significant reduction in growth, with a 28-day EC₁₀ (10% effect concentration) of 1.07% plastic weight in sediment dw.

Studies have also been conducted on coral species. Chapron et al. (2018) found that marine corals (Lophelia pertusa) exposed to LDPE microbeads had significantly lower prey capture rates and reduced skeletal growth rates and calcification compared to the controls. Hankins et al. (2018) found no significant effects on calcification in either the large polyp coral Montastraea cavernosa or the small polyp coral Orbicella faveolata despite active ingestion of PE microbeads.

Microplastic exposure has also been studied at early developmental stages for invertebrates. Lo and Chan (2018) found that larval and juvenile sea snails (Crepidula onyx) were not affected by exposure to environmentally-relevant concentrations of PS particles. At higher concentrations, the larvae grew slower and settled at a smaller size compared to control conditions. In addition, individuals exposed to microplastics only in their larval stages displayed slower growth rates even after the removal of the microparticles, indicating a possible legacy effect (Lo and Chan 2018). Similarly, blue mussel (Mytilus edulis) larvae with PS bead exposure experienced no changes in growth rate; however, there was an increase in the amount of abnormally developed larva (Rist et al. 2019). Beiras et al. (2018) found no significant effect on mussel embryonic development under static conditions from virgin PE microplastics. 

Primary producers

Green algae (Chlorella pyrenoidosa) exposed to PS beads displayed inhibited growth rates that corresponded to increasing plastic concentration (Mao et al. 2018). Reduced photosynthetic activity and damaged cell membranes were also evident; however, a recovery of algal biomass and photosynthetic activity was seen during the later phases of growth, which may be linked to detoxification mechanisms. Additionally, Gambardella et al. (2018) found that green microalga (Dunaliella tertiolecta) exposed to PS microbeads experienced a dose-dependent inhibition of growth: inhibition reached 40% at the highest concentration. 

Current studies also show an absence of significant effects on primary producers for endpoints tested. Sjollema et al. (2016) exposed both freshwater and marine microalgal species to uncharged virgin PS microbeads and negatively-charged beads and found an absence of significant effects on photosynthesis from exposure to all treatments. Further, Garrido et al. (2019) found no effect on the daily growth rate of the microalgae Isochrysis galbana exposed to PE particles at any of the tested concentrations. 

Soil

Experimental studies involving biota in the soil compartment are limited, but the studies that do exist show that microplastic exposure can negatively impact organism health and behaviour. 

Ju et al. (2019) showed that exposing soil springtails (Folsomia candida) to PE microplastics for 28 days led to an increase in avoidance behaviours and an inhibition of reproduction rate by up to 70.2% at the highest exposure concentration. Additionally, the exposed springtails had significantly decreased bacterial diversity in their guts. Similarly, Kim and An (2019) found that microplastic infiltration into soil system bio-pores caused movement inhibition in the invertebrate Lobella sokamensis. 

PS microplastics also caused toxicity to the soil invertebrate Caenorhabditis elegans following a three-day exposure period (Lei et al. 2018b). Nematodes exposed to 1.0 µm PS particles had lower survival rates, shorter average lifespans, decreased average body lengths, and significant damage to GABAergic neurons in comparison to the other microplastic sizes tested.

Sediment

Although the sediment compartment has also been less studied than the water compartment, the current literature indicates that microplastics may have adverse effects on sediment-dwelling organisms. 

Ziajahromi et al. (2018) exposed sediment-dwelling midge (Chironomus tepperi) larvae to four different size ranges of virgin PE microplastics to assess development. They concluded that midge survival was size-dependent; organisms exposed to microplastics that were similar in size to their normal food sources (10 to 27 µm) had a survival rate of 57% compared to 92% in the negative control group, as well as significantly smaller body sizes and head capsule lengths. Further, Leung and Chan (2018) found that PS microplastics significantly increased mortality and decreased body part regeneration in polychaetes (Perinereis aibuhitensis) after a four-week exposure period in a size-dependent manner. In addition, sediment-dwelling bivalves (Ennucula tenuis, Abra nitida) exposed to fragmented PE microplastics in three size classes displayed a dose-dependent decrease in energy reserves; however, no significant mortality was observed (Bour et al. 2018). The exposed E. tenuis also had significantly lower lipid content for only one condition, while lower protein content was observed in A. nitida from exposure to the largest particles at all concentrations. 

Nematodes (Caenorhabditis elegans) exposed to PA, PE, PP, and PVC microplastics had decreased survival rates, body length, and reproduction, as well as reduced calcium levels and increased expression of enzymes, indicating oxidative stress and intestinal damage (Lei et al. 2018a).

In contrast to the above-summarized research, the current literature also contains studies that show an absence of adverse effects on organisms exposed to microplastics in sediment. Redondo-Hasselerharm et al. (2018) observed no significant effects on survival or growth of the freshwater benthic macroinvertebrates Hyalella azteca, Asellus aquaticus, Sphaerium corneum, and Tubifex spp. from exposure to PS microplastics. Further, they observed no effects on the reproduction of the freshwater worm Lumbriculus variegatus.

6.2.3 Trophic transfer

There is limited information on the ability of microplastics to travel through different trophic levels, as seen in a food chain. Very few studies have looked at trophic transfer, and even fewer have studied the importance of bioconcentration, biomagnification, and bioaccumulation (Provencher et al. 2018b). Hammer et al. (2016) conducted one of the few studies that demonstrate vertical transfer of plastic particles within a food web. In that study, plastics found in the guts of great skuas (Stercorarius skua) from the Faroe Islands corresponded to the plastic contents of their prey (surface-feeding seabirds), implying indirect consumption. 

Additionally, Cuthbert et al. (2019) demonstrated transference of microplastics in predatory midge larvae (Chaoborus flavicans) that consumed mosquito (Culex pipiens) larvae exposed to 2 µm PS microplastics. They found that the amount of microplastics transferred correlated with feeding rates towards mosquito larvae. 

To study transfer along a natural food chain, Batel et al. (2016) exposed nauplii of the brine shrimp Artemia to microplastics ranging from 1 to 5 µm or from 10 to 20 µm, then fed the nauplii to zebrafish (Danio rerio). They observed that, while the zebrafish were able to uptake the microplastic particles, no significant accumulation or further retention was observed within their intestinal tract, and no transfer to other organs was observed. Similarly, Welden et al. (2018) found by examination of stomach contents that trophic transfer of microplastics occurred between sand eels (Ammodytes tobianus) and their predator, plaice (Pleuronectes platessa) from the Celtic Sea. However, the microplastics were egested in the plaice. 

Some studies suggest that unintentional ingestion, rather than trophic transfer, is the primary means by which microplastics are ingested. Chagnon et al. (2018) found no accumulation of microplastics in stomachs of yellowfin tuna (Thunnus albacares), a large predatory fish from Easter Island, despite plastics being found in the guts of its prey. Hipfner et al. (2018) also concluded that two fish species from the northeastern Pacific Ocean, the Pacific sand lance (Ammodytes personatus) and the Pacific herring (Clupea pallasii), do not act as significant conduits for the vertical transfer of microfibres to marine piscivores along the coast of British Columbia.  

6.2.4 Translocation

While mechanisms of translocation from an organism’s gut to other parts of its body are not well studied to date, the current literature has shown that translocation is usually size-dependent. For example, Lu et al. (2016) found that particles less than 5 μm can translocate to fish liver from the gut, while 20 µm particles cannot (Jovanović 2017). Smaller particles have the potential to more easily enter the circulatory system, but can also be egested more easily than larger microplastic particles (Jovanović 2017; Burns and Boxall 2018). 

Current studies show that translocation occurs in some organisms and organs, while other studies contradict these findings. For example, translocation of 0.5 μm PS spheres to the haemolymph, gills, and ovary was observed in crabs (Carcinus maenas) (Farrell and Nelson 2013). In zebrafish (Danio rerio), Lu et al. (2016) found 5 μm PS particles in the gills, liver, and gut, while 20 μm particles were only found in the gills and gut. In bivalves, tissue translocation of 3.0 or 9.6 μm PS spheres from the digestive tract to the circulatory system was seen in mussels (Mytilus edulis) by Browne et al. (2008). However, a study by Sussarellu et al. (2016) using Pacific oysters (Crassostrea gigas) showed no evidence of PS sphere (2 and 6 μm) translocation. Limited information in fish also shows very small amounts of microplastics in fish muscle (Karami et al. 2017a; Abbasi et al. 2018; Akhbarizadeh et al. 2018).

The conflicting results observed in these studies may be attributable to species-specific differences and/or false positive results that may occur as a result of leaching of fluorescent dye, which is often used to track particle ingestion. Schür et al. (2019) tested this theory and found that fluorescent droplets did not always co-localize with the plastic PS beads ingested by Daphnia magna. Using confocal laser scanning microscopy, 1 µm beads did not co-localize with the fluorescent dye in the gut and there was a rapid loss of fluorescence upon investigation. Fluorescence was also observed in lipid droplets outside of the digestive tract, but plastic particles were not detected in these same lipid droplets. Therefore, given that false positives may occur in uptake studies where precautions were not taken to avoid potential artifacts by ensuring the stability of dyes, controlling for dye leaching (e.g., by pre-washing the particles), or using microscopic imaging to confirm plastic presence, results should be interpreted with caution. 

7. Impacts on human health

7.1 Macroplastic

While people regularly observe and interact with macroplastics, human exposure to macroplastic pollution is not anticipated to be a concern. The effects of macroplastic pollution on human health are therefore not considered in this report.

7.2 Microplastic

Humans may be exposed to microplastics through the ingestion of food and drinking water (see Section 5.2) and the inhalation of indoor and outdoor air (see Section 5.1.3). The toxicity of microplastics via the ingestion and inhalation routes of exposure is reviewed below. Where possible, inferences are made from epidemiological studies on microplastics in humans and experimental studies on microplastics in animal models. A comprehensive review of in vitro studies on microplastics was not conducted as their relevance to human health is unclear. The effects of biofilms on human health are also discussed.

Upon ingestion or inhalation, microplastics may exert effects due either to their physical presence in the gut or lung or to the chemical composition of the plastic polymers themselves or their monomers, additives or sorbed substances. The World Health Organization (WHO) recently carried out an assessment of human exposure to microplastics in drinking water using conservative worst-case estimates of the levels of additives and sorbed chemicals on microplastics (WHO 2019). The Food and Agriculture Organization of the United Nations (FAO) and the European Food Safety Authority (EFSA) conducted a similar assessment of exposure to microplastics in seafood (EFSA 2016; FAO 2017). These evaluations concluded that exposure to microplastics and/or chemicals associated with microplastics are considered to be a low concern to human health (EFSA 2016; FAO 2017; WHO 2019). The reader is referred to those reports for further information on the exposure and risk assessments conducted therein.

7.2.1 Effects from oral exposure

Physicochemical properties affecting uptake and toxicity

Following ingestion, microplastic uptake and translocation are strongly dependent on the physicochemical properties of the ingested particles (FAO 2017; Wright and Kelly 2017; WHO 2019). Particle size is an important determinant of absorption through the intestinal epithelium. Smaller particles have larger surface-area-to-volume ratios, which can increase their ability to translocate to internal organs and increase bioreactivity (WHO 2019). A higher surface-area-to-volume ratio may also increase the sorption capacity of microplastics for environmental contaminants. Smaller particles may also be more susceptible to fragmentation, and while degradation of microplastics to smaller polymers has been demonstrated in the GI tract of Antarctic krill (Dawson et al. 2018), it is uncertain whether this occurs within the human GI tract (WHO 2019). Particle concentration can also influence toxicity, as higher concentrations are expected to overwhelm biological clearance mechanisms and cause responses that are otherwise not observed at lower doses (WHO 2019). At present, it is unclear how other properties, such as shape and surface chemistry, may affect the uptake, retention, and/or toxicity of ingested microplastics (Stock et al. 2019; WHO 2019).

Toxicokinetics

There are limited data regarding the fate of orally ingested microplastics in mammalian species. Available literature suggests that following oral ingestion, microplastics may remain confined to the GI tract, translocate from the GI tract into organs or tissues, and/or be excreted (EFSA 2016; FAO 2017). 

Several uptake mechanisms have been proposed for microplastics, including endocytosis via microfold cells (M cells) of the intestinal Peyer’s patches and paracellular persorption (see EFSA 2016, FAO 2017, and Wright and Kelly 2017 for an extensive review of the toxicokinetics of microplastics). Based on limited data, it is expected that the largest fraction of orally ingested microplastics (>90%) will be excreted in the feces (EFSA 2016; FAO 2017). Microplastics greater than 150 µm are also expected to remain confined to the gut lumen and be excreted, while only limited uptake is expected for smaller particles (EFSA 2016; FAO 2017; WHO 2019). Various types of microparticles have been shown to translocate across the mammalian GI tract into the lymphatic system at sizes ranging from 0.1 to 150 µm (Hussain et al. 2001; EFSA 2016; FAO 2017). For example, in one study, PVC microplastics (5 to 110 µm) were detected in the portal veins of dogs (Volkheimer 1975). Given these findings, it is possible that microplastics less than or equal to 150 µm may end up in the lymphatic system and result in systemic exposure, although absorption is expected to be low (≤0.3%; EFSA 2016; FAO 2017). Only very small microplastics (˂1.5 µm) are expected to enter into capillaries and penetrate deeply into tissues (Yoo et al. 2011; EFSA 2016). This is consistent with a recent 28-day study in which mice were administered high concentrations of a mixture of PS microplastics of various sizes by oral gavage three times per week (Stock et al. 2019). Only a few microplastics were detected in the intestinal walls (no quantitative analysis completed), representing a very low uptake by the GI tissue, and no microplastics were found in the liver, spleen or kidney. Conversely, another study reported significant translocation of 5 µm and 20 µm PS microplastics to the liver and kidney in mice (Deng et al. 2017), although these data are of questionable quality due to notable limitations in study design, data reporting, and biological plausibility of results (Tang 2017; Böhmert et al. 2019; Braeuning 2019). Based on a single human ex vivo placental perfusion model, fluorescently-labelled PS beads less than 240 nm may be taken up by the placenta (Wick et al. 2010).

Studies in humans

No epidemiological or controlled dose studies that evaluated the effects of orally ingested microplastics in humans were identified.

Studies in experimental animals

A small number of animal studies have evaluated the potential adverse effects of orally ingested microplastics (Merski et al. 2008; Mahler et al. 2012; Deng et al. 2017, 2018; Lu et al. 2018; Rafiee et al. 2018; Jin et al. 2019; Stock et al. 2019). Studies were limited to a few types of virgin microplastics and tested either unknown or high concentrations of microplastics that were not necessarily reflective of anticipated human exposure. Test concentrations in toxicity studies are orders of magnitude higher than would be anticipated for humans. Therefore, it was not possible to adequately evaluate the health risk of orally ingested microplastics with the currently available animal data (EFSA 2016; FAO 2017; Wright and Kelly 2017; WHO 2019). The WHO conducted the most recent review of the toxicological data on microplastics ingestion. Consistent with previous reviews by the EFSA (2016) and FAO (2017), the WHO concluded that there were insufficient data to allow for a robust evaluation of the potential human health risks of ingested microplastics, although there was no information to suggest it represented a potential human health concern (WHO 2019). Relevant toxicological studies are briefly summarized below, with more detailed descriptions, including test concentrations, provided in Table E-1 in Appendix E. 

In a 90-day study that was compliant with Organisation for Economic Co-operation and Development (OECD) test methods, rats fed a daily diet that contained up to 5% milled PE and PET fabric exhibited no treatment-related adverse effects on blood parameters, organ weights, or histopathology (Merski et al. 2008). Based on the absence of observed toxicity, the highest test dose was considered to be the no observed effect level (NOEL), equivalent to approximately 2 500 mg/kg body weight (bw)/day (WHO 2019). Fibre concentrations were not reported.

Other studies have reported adverse health effects in mice following the administration of very high oral doses of microplastics, several orders of magnitude above expected microplastic concentrations in food and drinking water (Deng et al. 2017, 2018; Lu et al. 2018; Jin et al. 2019). These studies have been extensively criticized for their lack of reliability and relevance (Böhmert et al. 2019; Braeuning 2019; Tang 2017; WHO 2019; Stock et al. 2019). Exposure to high concentrations of PS microplastics in drinking water was associated with alterations in lipid metabolism, gut microbiota composition, amino acid and bile acid metabolism, mucus secretion, and reduced intestinal barrier function in mice (Jin et al. 2019; Lu et al. 2018). Inflammation and lipid droplets were reported in the livers of mice administered high concentrations of PS microplastics by gavage (Deng et al. 2017), but the presence of these effects cannot be determined due to poor quality histological images (Braeuning 2019). Deng et al. (2017) also reported changes in metabolic profiles suggestive of disturbances in energy and lipid metabolism, oxidative stress, and neurotoxic responses. However, the relevance of these metabolic endpoints in assessing the potential human health effects of microplastics is difficult to interpret (Tang 2017; Braeuning 2019; WHO 2019). 

More recently, a 28-day mouse study evaluated the potential adverse effects of a mixture of various sizes of PS microplastics (1, 4 and 10 µm) administered via oral gavage three times per week using male heme oxygenase-1 reporter mice, a transgenic mouse model used to evaluate oxidative stress and inflammatory responses (Stock et al. 2019). In contrast to previous studies (Deng et al. 2017, 2018), the authors reported no evidence of oxidative stress or inflammation. While very high microplastic doses were selected for purposes of consistency with other rodent oral toxicity studies, the selected treatment scheme involved dosing the animals three times per week, which was intended to be more representative of a realistic human exposure scenario. However, given the high level of uncertainty surrounding human exposure to microplastics, it is unclear whether this experimental dosing regime (i.e., three times per week) was in fact more representative of human exposure than daily dosing regimes.  

7.2.2 Effects from inhalation

There may be hazards associated with the inhalation of microplastic particles due to their physical presence in the lung that are independent of chemical-related hazards. The scientific literature demonstrating the specific effects of microplastics on the lung is emerging, but their potential to cause effects in the respiratory tract or to translocate to other tissues remains uncertain. Still, inferences can be made from concepts of particle toxicology. Overall, toxicity related to the physical hazard of particles can include oxidative stress, cytotoxicity, inflammation, translocation to other tissues and, in some exceptionally elevated exposure circumstances, particle overload (elevated alveolar burden of particles that can impair clearance) (Prata 2018). Poorly soluble particles that are not inherently toxic, such as carbon black and TiO₂, have been shown to cause inflammation and tumours in rodents, albeit at very elevated levels of exposure (Borm and Driscoll 2019). Inhalation of fine particles is also associated with adverse respiratory and cardiovascular effects, although it is not possible to draw any conclusions regarding particle-driven effects of microplastics exposure at this time. 

The potential toxicity of particles will largely depend on particle size and shape, which will influence their deposition in the respiratory tract, their interaction with biological matrices, their potential to translocate, and the efficiency of particle clearance mechanisms. In general, inhalable particles larger than 10 µm in aerodynamic equivalent size will deposit mostly in the extrathoracic region, whereas particles below 10 µm can reach the tracheobronchial regions of the lung (US EPA 2009). It is expected that the majority of these particles will be removed from the airways by means of mucociliary clearance (i.e., trapping of the particles in mucus and coughing), though such clearance can result in ingestion of the particles and subsequent GI exposure (Gasperi et al. 2018). In theory, small particles below 2.5 µm in size can reach the alveolar region of the lung. These particles are removed through phagocytosis by alveolar macrophages, although there is some conflicting evidence demonstrating that very small particles in the nano-size range can evade alveolar clearance mechanisms and potentially accumulate in the lung, eventually reaching the interstitium (Li N et al. 2016). 

In the case of fibres, deposition patterns are more difficult to predict. Given their length, most microplastic fibres are expected to be deposited either in the extrathoracic region or in the upper airways and removed via mucociliary clearance (Gasperi et al. 2018). In general, longer plastic fibres, although flexible, are more likely to be associated with evasion of clearance mechanisms (Prata 2018). The area in which deposition occurs and residency time in the lung will greatly influence physical hazards associated with microfibres. Although there are insufficient data specific to microplastics, the observation of plastic microfibres in lung tissue biopsies of workers from a synthetic textile industry, as well as in healthy and neoplastic lung tissues from lung cancer patients, substantiates the plausibility of pulmonary microplastic retention through inhalation (Pauly et al. 1998; Wright and Kelly 2017). 

There are few studies that evaluate microplastic particle translocation from the lung following inhalation. It is possible that microplastics can translocate from the lung to systemic circulation or to the lymphatic system, potentially reaching other tissues. One study has examined translocation following intratracheal instillation in pregnant rats (equivalent to 2.4 x 10¹³ particles) and revealed systemic translocation to placenta, whole pup, fetal liver, heart and spleen (Fournier et al. 2018). Rats intratracheally instilled with radiolabelled PS particles of 56.4 and 202 nm in size exhibited only a small fraction (<2.5%) of particle translocation into systemic circulation in healthy rats, which increased to 4.7% for the smaller particles in the presence of lipopolysaccharide-induced lung inflammation (Chen et al. 2006). The likelihood of translocation is expected to increase with decreasing particle size and increased residency time as well as in individuals with compromised lung function and during events of inflammation (i.e., due to increased cellular permeability) (Galloway 2015). The alveolar region of the lung is a site of potential concern, in part because smaller particles can penetrate this region of the lung (and because they are, by nature, more reactive due to their high surface area), and in part because of the exchanges with systemic circulation that occur at this site. In the upper airway, particles may diffuse through mucus and reach underlying epithelium, where translocation may occur. However, diffusion through mucus is not expected to occur for insoluble particles such as microplastics. It should be noted that, in rats, ultrafine particles have been shown to reach brain tissue via translocation from the nasal cavity through the olfactory nerve (Oberdörster et al. 2004). 

There is a paucity of information on the physical hazards related to inhalation of microplastics. Future studies should focus on confirming and exploring the toxicological mechanisms of the physical hazards associated with microplastics, including their effects on the lung and cardiovascular system and their capacity to translocate to extra-pulmonary tissues.  

Studies in humans

In the only controlled dosing studies of microplastics in humans, participants were exposed to printer toner, which was not considered relevant for this evaluation. Epidemiology studies of microplastics in indoor or ambient air could not be found for the general population. Studies on the health effects of microplastics are limited to several occupational epidemiology studies and a lung biopsy study; these studies are summarized below. 

Two reviews (Wright and Kelly 2017; Prata 2018) summarized the outcomes of occupational epidemiology studies in individuals who worked with synthetic textiles, nylon flock, and PVC. The studies identified associations between work in these industries and increases in adverse respiratory effects, including airway lesions and fibrosis, decreased pulmonary function, wheezing, dyspnea, inspiratory crackles, chronic cough, chronic mucous production, eye and throat irritation, increased bronchial responsiveness, bronchitis, bronchiolitis, emphysema, asthma, pneumoconiosis, interstitial lung disease, foreign body granulomas, and acute respiratory failure (Wright and Kelly 2017; Prata 2018). Several studies also found associations between work in these industries and cancers of the digestive system and respiratory tract, but not all studies investigating these effects identified the association. Despite the associations between exposures to plastic particulates or fibres and adverse health effects, no firm conclusions on human health effects can be made owing to confounding variables such as co-exposures with other workplace hazards that could contribute to respiratory effects.

A third review discussed the epidemiological evidence of health effects in women working in plastics manufacturing and processing industries, but did not specifically address microplastics (DeMatteo et al. 2012). Epidemiology studies identified associations between work in plastics industries and breast cancer, spontaneous abortion, and infertility. As exposures to microplastics were not specifically discussed in these studies, it is unclear whether associations with these health effects are related to inhalation of plastic particulates and fibres or exposure to other substances used in the production of plastic.

Epidemiology studies have been developed for other occupations with exposure to microplastics. However, most studies limited exposure categorization to occupation, and therefore adverse outcomes from exposure to microplastics were not specifically investigated. A small subset of the epidemiology studies included analyses specifically related to exposure to plastic particulates or fibres; only these studies are discussed here. No increases in lung or respiratory tract cancer were associated with exposure to PU dust in polyurethane foam (PUF) workers (Sorahan and Pope 1993; Mikoczy et al. 2004; Pinkerton et al. 2016). In pattern and model makers, an increase in lymphocytopenia was significantly associated with exposure to plastic dusts, but no exposure–response relationship was observed (Demers et al. 1994).

The relevance of occupational data on airborne microplastics to the general population is unknown, as extrapolation from high-dose occupational exposures to lower doses, as would be expected for the general population, is difficult in the absence of health effect data at lower concentrations. A further limitation of the dataset is that most studies did not investigate the impact of dose-response on the health outcomes. Additionally, workers in the studies might have had co-exposures to other chemicals associated with adverse health effects, such as monomers, catalysts, additives, and other compounds used in the workplace.

Studies in experimental animals

Studies of inhaled microplastics were identified for rats (Laskin et al. 1972; Thyssen et al. 1978; Hesterberg et al. 1992; Warheit et al. 2003; Ma-Hock et al. 2012), hamsters (Laskin et al. 1972), and guinea pigs (Pimentel et al. 1975). The microplastic constituents in the studies included PP fibres (Hesterberg et al. 1992), PU particulate (Laskin et al. 1972; Thyssen et al. 1978), nylon fibres or particulate (Pimentel et al. 1975; Warheit et al. 2003), PAN particulate (Pimentel et al. 1975), and acrylic ester copolymer (Ma-Hock et al. 2012). Exposure duration varied, with one longer-duration study of 325 days, two studies of subchronic duration (12 to 13 weeks), and three studies of subacute duration (5 to 30 exposure days). Detailed descriptions of these studies, including test concentrations and results, are presented in Table E-2 in Appendix E.

Effects observed in inhalation studies tended to be consistent and independent of duration, type of plastic, and species. Observations consistent with foreign body reactions were common in the studies. This included an increase in activity or number of inflammatory cells, which contained fibres or particles (primarily in lung tissues and bronchoalveolar lavage fluid [BALF], but also in the lymphatic system) and which were often accompanied by granulomas. In areas of lungs associated with particle deposition, hyperplasia, emphysema, and edema were observed. Studies in which animals were euthanized at various timepoints post-exposure tended to indicate a reversibility of effects, suggesting that effects are adaptive rather than adverse responses. No dose-related effects were observed in mortality, survival time, behaviour, clinical observations, tumour incidence, or fibrosis. LOECs adjusted to reflect intermittent exposure ranged from 0.48 to 2.3 mg/m³. One exception was for the shortest duration study, in which no treatment-related changes in BALF or histology were observed up to the adjusted no observed effect concentration (NOEC) of 2.7 mg/m³ in rats exposed for 5 days and followed up to 24 days post-exposure (Ma-Hock et al. 2012). However, most studies were not performed according to OECD test guideline methods. Moreover, the human relevance of these animal studies is unclear, as exposures in the studies are much higher than would be expected in humans under typical exposure scenarios.

Inhalation studies are also supported by observations in intratracheal instillation studies in rats. Exposures in the studies were to PVC particulate (Agarwal et al. 1978; Pigott and Ishmael 1979; Xu et al. 2004), nylon fibres or particulates (Porter et al. 1999), PS particulate (Brown et al. 2001; Fournier et al. 2018), or PU particulate (Stemmer et al. 1975). Most of the studies incorporated only one exposure level, and contained a single intratracheal instillation, except for one group in Fournier et al. 2018 (every second day). The rats in the various studies were followed from 1 day to 24 months post-instillation. In general, the foreign body reactions observed in inhalation studies were also observed in the intratracheal studies. One study demonstrated that effects from washed PVC particulates were equal to or greater than those from unwashed PVC particulates, suggesting that adverse effects were from the plastic particulate itself rather than from adsorbed additives (Xu et al. 2004). Additional pulmonary effect observations are outlined in Table E-2 in Appendix E. A developmental study also observed an increase in fetal reabsorption sites and evidence of particle translocation from the lungs (placenta, whole pup, fetal liver and heart, and maternal heart and spleen) (Fournier et al. 2018). Although results from the intratracheal studies corroborate effects observed in the inhalation studies, they should not be primarily relied upon for quantitative dose-response assessments because the route of exposure does not accurately represent deposition patterns and dosing that would be observed from inhalation.

A review of the toxicology of p-aramid (an aromatic PA commonly known as Kevlar) fibrils was also identified (Donaldson 2009). Studies of rat lungs identified effects at high exposure levels, such as inflammation, increased cell proliferation, fibrosis, and development of cystic keratinizing squamous cell carcinoma (a tumour stated to be of questionable relevance to humans due to an absence of a human homologue). 

7.2.3 Effects of biofilms

Microplastics provide a unique and extensive surface for microorganisms to attach to and colonize in water environments, forming biofilms (Zettler et al. 2013; De Tender et al. 2015; McCormick et al. 2016; Oberbeckmann et al. 2018; Kettner 2018; Arias-Andres et al. 2018, 2019). However, very few studies have analyzed microplastic-associated biofilms. 

Biofilms consist of accumulations of microorganisms, typically encased in a self-secreted matrix of extracellular polymeric substances, containing both organic and inorganic matter (Liu et al. 2016; Prest et al. 2016; WRF 2017). The structure of the extracellular polymeric substances provides protection from stressors (e.g., predators, disinfectants), and aids in uptake and utilization of nutrients (Flemming and Wingender 2010; Prest et al. 2016). Biofilms are ubiquitous in the environment (Hall-Stoodley et al. 2004; Yadav 2017) and in drinking water distribution systems (Liu et al. 2016; Prest et al. 2016; WRF 2017), where they provide a habitat for the survival and growth of microorganisms, including potential pathogens (US EPA 2002; Batté et al. 2003; Berry et al. 2006; Liu et al. 2016).

The higher surface-to-volume ratio of microplastics facilitates the absorption of organic matter, which serves as nutrients for microorganisms, thereby promoting biofilm formation. The transport of microplastics over long distances and through the water column (Peng et al. 2017) affords opportunities for attachment of microbial “hitchhikers” and biofilm formation (Kirstein et al. 2016; Zalasiewicz et al. 2016; Keswani et al. 2016). These plastic-associated biofilm communities are sometimes referred to as “plastispheres” (Zettler et al. 2013) and tend to differ from microorganisms in surrounding water or on natural aggregates/particles (Zettler et al. 2013; Hoellein et al. 2014; McCormick et al. 2016; Oberbeckmann et al. 2016; Kettner et al. 2017; Arias-Andres et al. 2018, 2019). Gene sequencing studies have demonstrated that microbial communities on microplastics are less diverse than those on non-plastic substrates (Zettler et al. 2013; Harrison et al. 2014; McCormick et al. 2014, 2016; Ogonowski et al. 2018a), suggesting that microplastics may select for specific microbial colonizers. In other words, the physicochemical properties of microplastics influence the composition and structure of the associated biofilm community (Bhardwaj et al. 2013; Zettler et al. 2013; Harrison et al. 2014; McCormick et al. 2014, 2016). It is unclear what impact this has, but some have hypothesized that it may result in reduced competition and predation, leading to the emergence of potential pathogens (Amalfitano et al. 2014; Keswani et al. 2016; Andrady 2017). Other factors, including environmental conditions (e.g., salinity, temperature), can also influence biofilm formation on microplastics (Harrison et al. 2018; Oberbeckmann et al. 2018; WHO 2019). In addition, microorganism features, such as the hydrophobicity of their cell walls and cell surface charge, can impact attachment to microplastics (Rummel et al. 2017).

Biofilm constituents commonly found on microplastics include various non-pathogenic microorganisms, comprising species of Pseudomonas, Arcobacter, Erythrobacter, Streptococcus, Staphylococcus, Aspergillus, Penicillium and Phanerochaete (Bhardwaj et al. 2013; McCormick et al. 2014). Pathogenic bacterial sequences, primarily those of Vibrio, have been detected in microplastic-associated biofilms (Zettler et al. 2013; De Tender et al. 2015; Kirstein et al. 2016). However, aside from one study (Kirstein et al. 2016), species identification was not possible, and it is therefore unknown whether the organisms were of human health concern. In the study, Vibrio spp. of potential human health significance were identified, namely V. parahaemolyticus, V. fluvialis, and V. alginolyticus.

The increased cell density and proximity, improved nutrient availability, and protection afforded by an extracellular polymeric substances matrix make biofilms an ideal environment for interactions between microorganisms, including those on microplastics. Among these interactions is conjugation, the transfer of genetic material through direct cell-to-cell contact (Cook et al. 2011; Stalder and Top 2016). Conjugation is a method of horizontal gene transfer (HGT), the primary mechanism for the spread of antibiotic resistance, whereby a mobile genetic element (MGE), such as a plasmid, containing antibiotic resistance genes (ARGs), is transferred from a donor to a recipient cell (Von Wintersdorff et al. 2016). A few studies have shown that ARGs are more frequently transferred between microplastic-associated biofilm members than free-living bacteria or biofilms associated with natural aggregates (Arias-Andres et al. 2018; Eckert et al. 2018a,b; Imran et al. 2019; Laganà et al. 2019). Transfer also occurred between a broader (i.e., more distantly related) group of microorganisms on the microplastics than in the natural environment. These findings suggest that microplastic-associated biofilms provide a favourable environment (i.e., “hot spot”) for HGT events and may select for antibiotic resistant microorganisms and ARGs, which may then be transported to different habitats. Transfer of ARGs via microplastics has been observed between wastewaters and the aquatic environment (Eckert et al. 2018a,b). Transfer events on microplastics may be further amplified through exposure to metals, as metal resistance genes are present on the same plasmid as antibiotic resistance genes (Baker-Austin et al. 2006; Wright et al. 2006; Seiler and Berendonk 2012; Zhang et al. 2018; Imran et al. 2019). 

Although research in this area is very limited, studies suggest that plastic-associated biofilms in water may harbour potential human pathogens and ARGs. Given that microplastics can travel long distances (see Section 4, WHO 2019), there is a possibility that these organisms and/or ARGs may be dispersed across waters and enter drinking water sources. Despite this, there is no indication of how prevalent these organisms are or of how long they persist and/or remain infectious while in a plastisphere. Moreover, conventional drinking water treatment is expected to significantly reduce microplastics and inactivate associated biofilm organisms (see Section 4.1.3, WHO 2019). Thus, there is currently no evidence to suggest that microplastic-associated biofilms in drinking water pose a risk to human health.

Microorganisms might also adhere to the surface of airborne microplastics, but data are limited. Microorganisms have been measured in airborne particulates (Noble et al. 1963; Brodie et al. 2007), although no data exist specifically for plastic particulates. Adherence and growth of microorganisms on airborne microplastics might be limited because they could be dependent on the contact of microorganisms and microplastics in the environment. However, if contact does occur, the plastic particulates might protect and shield adhered microorganisms (Prata 2018). While no data could be found on the characterization of microbial communities potentially colonizing airborne microplastics, lung infections could theoretically occur if pathogenic species were adhered to microplastics and inhaled (Prata 2018).

8. Transport of chemicals

In addition to the physical hazards presented by plastic particles themselves, it is possible that effects could occur as a result of exposure to residual monomers, chemical additives, and sorbed environmental contaminants (e.g., persistent organic pollutants [POPs] and metals) that may leach from microplastic particles (Munier and Bendell 2018; SAPEA 2019). Although there is potential for environmental or human exposure to these compounds, these chemicals are considered to be under the purview of various programs at ECCC and Health Canada and will continue to be managed in accordance with those programs. 

Any effects observed from the transport of chemicals are highly context dependent. For example, the type of plastic and the physicochemical properties of the sorbed chemical are known to have an effect on sorption ability. In general, PE shows a greater ability to sorb contaminants, while PET and PVC have a lower sorption capacity (Alimi et al. 2018). Plastics with high surface-area-to-volume ratios (i.e., small, elongated, or have an irregular shape) tend to have higher sorption capacities (Rochman 2015). For instance, PVC was shown to have significantly greater absorption of copper than PS, which could be due to its greater surface area and polarity (Brennecke et al. 2016; Munier and Bendell 2018). Sorption can also be affected by factors such as age, shape, molecular weight and porosity of the particle, temperature, salinity and pH of the environment (increased salinity and particle age tend to increase sorption, and alkaline environments favour sorption of cations), and the concentration of metals and other contaminants in the surrounding waters (Rochman 2015; Alimi et al. 2018; Munier and Bendell 2018; Guo and Wang 2019a). Di and Wang (2018) sampled surface waters and sediments from China’s Three Gorges Reservoir and found that several contaminants were adsorbed to the surface of the recovered microplastics, including organic solvents and pharmaceutical intermediates.

The properties of the receiving environment can also affect contaminant transfer. Mohamed Nor and Koelmans (2019) found that the transfer of polychlorinated biphenyls (PCBs) from microplastics in simulated gut fluid is biphasic and fully reversible. More specifically, the effect of microplastics in the gut depends on the contents of the gut system. Ingested plastics acted as a source of hydrophobic organic compounds (HOCs) in clean gut systems, whereas in contaminated gut systems, clean microplastics rapidly extracted PCBs from food or other organic matter (Mohamed Nor and Koelmans 2019). The authors concluded that chemical contamination and cleaning can occur simultaneously when microplastics are ingested. 

Although many of the compounds associated with plastic have short biological half-lives and are not persistent, plastic particles within the body could present a long-term source of exposure to the chemicals (Engler 2012). While recent reviews indicate that there is a low health concern for human exposure to chemicals from ingestion of microplastics from food or drinking water (EFSA 2016; FAO 2017; WHO 2019), further research would be required before a human health risk assessment on microplastics is possible. No data could be found on the transfer of these compounds in the human respiratory or GI tract.

Sorbed chemicals

Provencher et al. (2018c) found no significant correlations between concentrations of various PCB congeners in Northern fulmars (Fulmarus glacialis) and the amount of ingested plastics when using a toxic equivalency factor (TEF) approach. They found that plastics did not contribute to the PCB concentrations in the birds and that the PCB congener profile between ingested plastics and the liver differed (Provencher et al. 2018c). This could be the result of the ability of Northern fulmars to metabolize or bio-transform contaminants such as PCBs (Letcher et al. 2010; Provencher et al. 2018c). In a study using goldfish (Carassius auratus), Grigorakis and Drouillard (2018) observed lower dietary assimilation efficiencies (13.4%) for PCBs sorbed to microplastics compared to efficiencies (51.6%) for PCBs associated with food. The authors concluded that the lower bioavailability of PCBs associated with microplastics indicates that microplastic presence is unlikely to increase PCB bioaccumulation in fish. In a study by Devriese et al. (2017), Norway lobsters (Nephrops norvegicus) exposed to PCB-loaded PE or PS microplastics showed no significant bioaccumulation of the chemicals, with uptake of the PCBs being limited. Furthermore, Gerdes et al. (2019) found a positive correlation between the elimination rate of PCBs in Daphnia magna and the presence of microplastics. More specifically, the presence of microplastics together with PCBs was able to increase the elimination rate of high-molecular-weight PCB congeners in D. magna fourfold.

Diepens and Koelmans (2018) introduced a theoretical model simulating the transfer of microplastics and HOCs in aquatic Arctic food webs. Simulated scenarios showed that PCBs biomagnify to a lesser extent with higher levels of microplastic ingestion, which supports the evidence previously described. Conversely, the same model also indicated that polycyclic aromatic hydrocarbons (PAHs) biomagnify more with elevated levels of microplastic ingestion. Under different conditions, Magara et al. (2018) found that the uptake and accumulation of fluoranthene (a PAH) in blue mussels (Mytilus edulis) were not affected by incubation with microplastics and that incubation with microplastics reduced the bioavailability of fluoranthene. In a study modelling the transfer of POPs from PVC and PE to benthic invertebrates, fish, and seabirds, Bakir et al. (2016) found that food and water were the main pathways of exposure for all organisms, and input from microplastic particles was negligible.

Tanaka et al. (2013) studied the occurrence of polybrominated diphenyl ethers (PBDEs) in tissues of short-tailed shearwater (Puffinus tenuirostris) seabirds, in their natural prey, and in plastics in the stomachs of the seabirds. In three of the 12 short-tailed shearwaters examined, they detected higher-brominated congeners of PBDEs that were not present in their prey (i.e., lanternfish and squid), which were also sampled from the same area as the seabirds. However, they did detect these PBDEs on the plastics found in the stomachs of the three birds, which suggests that plastic-derived chemicals were transferred from the ingested plastic to the seabird tissue.

Hydrophobic POPs of potential human health concern (such as PCBs, PAHs, and organochlorine pesticides) can readily sorb to plastics. For that reason, plastic compounds such as PE and PU are used as passive samplers in environmental monitoring (WHO 2019). 

Studies on microplastic-associated sorbed pollutants in drinking water could not be identified, but increased POPs in microplastics have been measured in marine environments and shorelines near urban environments (Wang et al. 2017; Pellet Watch 2019). 

Limited data exist on the sorption of chemicals to microplastic particulates in outdoor air, indoor air, or indoor dust. Adsorption of organic pollutants in air to plastic particulates could theoretically occur, but would be dependent on the duration of microplastic suspension in air (Prata 2018). One study reported that no significant adsorption of PCBs, dichlorodiphenyldichloroethylene (DDE) or nonylphenol occurred on virgin PP pellets released to the atmosphere for six days (Mato et al. 2001). Therefore, the contribution of microplastics to inhalation of sorbed chemicals is unknown but potentially limited, although it is anticipated to be dependent on environment (e.g., urban versus rural environments, proximity to point sources). Overall, current research shows that, while microplastics are able to transport POPs, the evidence suggests that the impact of this exposure pathway is minimal (Burns and Boxall 2018).

Monomers

Plastics are manufactured through the polymerization of monomers, which vary in toxicity. Some of the more hazardous compounds include acrylonitrile, acrylamide, 1,3-butadiene, ethylene oxide, and vinyl chloride (Lithner et al. 2011). Depending on the polymerization process, the plastic material can contain a range of concentrations of residual monomers (from negligible amounts to up to 4%) due to incomplete polymerization (Araújo et al. 2002; Lithner et al. 2011). Plastics can also be degraded (through biological processes and weathering) into monomers and oligomers, but few data exist on the contribution of these processes to human exposures to monomers (WHO 2019).

Additives

As discussed in Section 2, plastic additives can include polymer stabilizers, flame retardants, lubricants, plasticizers, and colourants. Compounds with potential human health effects that are additives of plastics include phthalates, PBDEs, lead, and cadmium (WHO 2019), among others. Plastic additives are mostly not co-polymerized, resulting in increased likelihood of being leached into the environment (Wright and Kelly 2017; Hahladakis et al. 2018). Molecular weight of additives and age of plastics are factors that can influence the rate of migration of additives from plastics to the surrounding environment (Hansen et al. 2013; Suhrhoff and Scholz-Böttcher 2016; Jahnke et al. 2017). Limited data exist on the contribution of microplastics to concentrations of plastic additive compounds in the environment, but there is evidence of potential migration pathways for the compounds in sources of human relevance, such as food (Helmroth et al. 2002; Muncke 2011), water (WHO 2019) and indoor dust (Rauert et al. 2014).

9. Knowledge gaps and considerations for future research

Several knowledge gaps were identified during the writing of this report and are outlined below with the objective of encouraging further research. Addressing these knowledge gaps will contribute to the understanding of the environmental and human health risks of plastic pollution and will inform science-based policy and regulatory decisions related to plastic pollution.

9.1 Occurrence

While the approach to observing macroplastics is relatively obvious given their size, there is a general lack of consistency and reliability in the methods used to sample and quantify microplastics in the environment and other media (e.g., drinking water and food). Many studies rely only on visual identification to determine if a particle is plastic. This can lead to a high false positive rate (especially at sizes smaller than 1 mm) and does not allow for proper characterization of plastics. For instance, when fibres visually identified as microplastics from the GI tracts of eelpout (Zoarces viviparus) were analyzed with μATR-FTIR by Wesch et al. (2016), none of the fibres were determined to be of synthetic origin. Given these findings, the authors question whether visual identification alone is sufficient to determine if microfibres are microplastics and call for standardized approaches for identifying and monitoring microplastics. Non-specific fluorescence staining methods have been suggested as a potential rapid-screening approach for detecting and quantifying microplastics in various media (Erni-Cassola et al. 2017; Maes et al. 2017; Prata et al. 2019). However, a major drawback of these staining methods is the possible introduction of false positives through the staining of biological organisms, such as marine algae or organic matter. 

Spectroscopic techniques, such as FTIR, Raman spectroscopy and pyrolysis GCMS, are currently the preferred methods for plastic characterization and are often used following separation of suspected plastic particles from sample media and visual identification using a microscope. Although they increase the accuracy of the identification of microplastics, spectroscopic analyses have limitations that can lead to the underestimation of microplastics in samples. With Raman spectroscopy, the generation of fluorescence can overpower the Raman spectrum produced, which can hinder the identification of potential plastics (Rezania et al. 2018). Furthermore, the signal can be heavily influenced by dyes, as well as by microbiological, organic and inorganic substances (Nguyen et al. 2019). With infrared (IR) spectroscopy, black or dark particles are not detected because they have a high absorption rate (Rezania et al. 2018), and particles below 20 µm may not yield enough absorbance interpretable spectra (Li et al. 2018b). Pyrolysis GCMS lacks reproducibility, as results are highly dependent on sample preparation and pyrolysis type. Thermal desorption GCMS is best used for samples of high mass (up to 100 mg) but lacks the sensitivity of pyrolysis GCMS (Nguyen et al. 2019). Microplastic counts can also be overestimated. Using SEM with energy dispersive X-ray spectroscopy (EDS), Anderson et al. (2017) found that, on average, 23% of the particles that were visually identified as plastics were not plastic. Burns and Boxall (2018) highlight that the error rate for visually identifying particles as plastic ranges from 33% to 70%. While analytical methods may help to confirm the synthetic nature of microplastics sampled in the environment, the inconsistencies in sampling methods (e.g., size of subsamples and sampling strategies) can limit the comparability of such analyses.

Studies investigating the occurrence of plastics in the environment and other media often use different units to report plastic abundance (e.g., plastics per area vs. plastics per unit volume), thereby limiting comparisons between studies and the generalizability of results. Standardized reporting metrics are required to ensure reporting consistency and study comparability (Burns and Boxall 2018). Another major gap in the analytical process is that there are no inter-laboratory studies, which are useful for method validation. Furthermore, due to variability and difficulty in quantifying microplastics, large standard deviations have been reported for the occurrence of microplastics in the environment and, in some instances, the standard deviation value exceeds the reported measurement. 

In water, microplastics are sampled at size ranges that are compatible with available sampling apparatus (e.g., trawl nets, which have a mesh size of 300 to 350 μm). This means that microplastics smaller than 300 µm can often go undetected. This is an issue for microfibres in particular given their narrow size (Covernton et al. 2019). Sampling methods therefore need to be developed to support the characterization of the smaller size fractions of plastics in the environment. Further, sampling depths vary across studies and are not standardized (i.e., trawl nets would be biased to less dense plastics that are present near the surface of surface waters, and studies conducted at a greater depth would be biased against denser plastics). 

A limited number of published studies report on the environmental monitoring and effects of microplastics in freshwater and terrestrial environments (Burns and Boxall 2018; Provencher et al. 2018b). There is a need to expand work to include monitoring studies to other ecosystems, particularly terrestrial ecosystems. 

In terrestrial matrices, studies of microplastic occurrence are scarce, possibly due to difficulty in translating research ideas in a marine context to a terrestrial context (Rillig 2012; da Costa et al. 2019). For example, there are no parallels for the accumulation of microplastics along shorelines in a terrestrial setting. In addition, it is often more difficult to isolate and characterize microplastics from a soil matrix; soil can contain varying levels of organic matter, which can distort signals and present problems when using FTIR and Raman spectroscopy for plastic characterization (Bläsing and Amelung 2018). Furthermore, there is a lack of standardized protocols for soil sampling and analysis in various soil types (da Costa et al. 2019). It has been suggested that a standard step-by-step approach be employed for terrestrial samples, involving removal of adherent fragments, mineral phase, and organic matter, followed by microplastic identification and quantification (da Costa et al. 2019).

There is also a lack of appropriate quantitative data for microplastic presence in drinking water and in water discharged after wastewater treatment, and limited information is available on the fate of microplastics during the wastewater treatment process, including particle breakdown, particle composition, removal efficiency, and subsequent release of these microplastics to other environmental compartments. 

Occurrence data for microplastics in food is also scarce, with little to no Canadian-specific data. Data that do exist are focused on wild marine fish and shellfish, with limited occurrence data for freshwater and farmed species or other foods. In addition, occurrence data are needed for the tissues and organs of animals that are consumed by humans. Data are lacking on the potential effects of cooking or food processing (e.g., fresh versus frozen food) on microplastic concentrations, the impact of the food matrix on microplastic bioavailability (e.g., water-based versus solid/dry foods), and the potential point source(s) of exposure to microplastics in food. Further studies are needed to determine whether food manufacturing, processing and/or handling as well as food packaging materials may contribute to microplastic concentrations in food.

There are currently no validated or recognized methods for the collection or analysis of microplastic samples in air, and little information is available on the partitioning of microplastic particles between air and dust. In order to accurately assess microplastic exposure from air, there is a need to develop and validate accurate, precise and reproducible methods for the sampling, extraction, characterization, and quantification of airborne microplastics and microplastics in settled dust and air, including robust quality assurance and control protocols. As Canadians spend approximately 90% of their time indoors, data on both indoor and outdoor microplastic exposures are needed to determine personal exposures, to understand their sources, pathways, fate, and distribution, and to identify and prioritize specific microplastic categories or mixtures for future research. There is also a need to explore the relationship between airborne microplastic particles and particulate matter. For example, knowing what proportion of particulate matter is composed of plastic polymers and knowing whether airborne plastic particles behave similarly to other airborne particulates would be useful in determining whether inferences can be made from the wealth of knowledge that exists on particulate matter.

Several researchers have identified the need for standardized protocols and stricter quality assurance in literature to ensure the availability of more high quality occurrence and exposure data in all media (Burns and Boxall 2018; Hermsen et al. 2018; Gouin et al. 2019; Koelmans et al. 2019). For sampling methods, this would include collection media, equipment, and handling procedures, as well as laboratory analysis practices. Due to the ubiquity of plastics, additional care must be taken throughout the entire process, from sample collection to laboratory analysis, to prevent sample contamination.

The importance of protocol development can be demonstrated by Provencher et al. (2017, 2019), who developed the only standardized protocols for monitoring and studying ingested plastics in seabirds. They include standardized field and lab techniques, as well as reporting guidelines for data (Provencher et al. 2017, 2019). The use of these standardized techniques by the international seabird community has led to spatial and temporal tracking of trends in plastics in the marine environment. 

It has been recommended that standardized quality criteria be developed that can be used to evaluate the appropriateness of studies on microplastic occurrence and effects. Hermsen et al. (2018) proposed several areas that should be evaluated when scoring the quality of microplastic ingestion studies: sampling method and strategy, sample size, sample processing and storage, laboratory preparation, clean air conditions, negative and positive controls, target component, sample treatment, and polymer identification. When reviewing current studies on microplastic ingestion by biota, they identified negative controls, polymer identification, laboratory preparation, and sample treatment as areas that were particularly lacking in quality and available information. Koelmans et al. (2019) evaluated 50 microplastic studies in freshwater surface water and drinking water using the same method identified by Hermsen et al. (2018). Only four studies scored positively on all proposed quality criteria; 92% of the reviewed studies were not considered complete or reliable on at least one criterion. It should be noted that Hermsen et al. (2018) and Koelmans et al. (2019) acknowledge that their criteria are not an absolute judgement of the value of studies since not all aspects of studies could be captured in their scoring system. Moving forward, the use of standardized quality criteria will ensure that only data of acceptable quality are being used to inform scientists and policy makers and that the data are both reproducible and directly comparable. 

There is also a paucity of data on the common or important sources of microplastics in the environment and other media, such that identifying source contributions of microplastics is difficult. There is a need to develop libraries that can be used to link samples to their sources using their chemical composition (polymer and additive chemicals) and other physical properties. Furthermore, establishing a taxonomy of microplastics based on morphology may also be informative in determining sources (Helm 2017).

Lastly, data on the occurrence and effects of nanoplastics are still emerging and poorly understood. It is unclear whether and how nanoplastics may form in the environment (e.g., whether they are formed by processes such as the weathering of macroplastics or microplastics). There is a lack of appropriate analytical methodologies for nanoscale materials in all media, making accurate measurements of environmental occurrence and behaviour of nanoplastics difficult to evaluate (SAPEA 2019). As nanoplastics are inherently more difficult to test and measure, the importance of plastics fragmenting to the nanoscale remains unclear at this time (Koelmans et al. 2015). 

9.2 Environmental effects

The size ranges and concentrations of microplastics used in ecotoxicological research do not reflect the concentrations or sizes of microplastics collected in the environment using current sampling techniques. Microplastic effects studies are often performed using concentrations that are much higher than those currently reported in the environment or very small microplastics for which limited occurrence data exists (SAPEA 2019). Researchers studying effects should use plastics of similar size, shape, and composition to those found in the environment. Additionally, there is a need to further investigate the relationship between microplastics and natural particles that exist in the environment that induce similar effects in biota. Currently, experimental designs do not differentiate plastic-specific effects from those caused by other particles, such as clay or cellulose (Ogonowski et al. 2018b). Furthermore, effects studies are largely conducted with PS microplastic spheres, which are not representative of plastics found in the environment. More frequently detected microplastics (i.e., PP, polyester, and PA, among others) are underrepresented in effect studies (SAPEA 2019). 

There is therefore a need to develop standard methods for testing the potential for adverse effects associated with exposure to plastic. For example, there is a need to evaluate the relationship between the properties of plastic (e.g., particle size, polymer type, shape and particle number) and toxicity. There is a corresponding lack of consistency in reporting test concentrations in studies; some studies report weight/volume, while others report particle number/volume. 

Often, microplastics used in toxicity testing are purchased. These microplastics would not be ideal models for microplastics that would be encountered in the environment as they can contain additives such as surfactants. For example, Pikuda et al. (2019) found that the acute toxicity to D. magna was associated with sodium azide, a surfactant, and not the plastic PS particles themselves. When the sodium azide was removed from the plastics, the PS particles no longer caused mortality. Thus, plastic particles used in toxicological studies should be washed to remove any additives that may cause effects that can be confused with effects caused by the particles themselves. Currently, this is not standard practice, and was not considered when selecting studies for this report.

Burns and Boxall (2018) suggested that research in this field should move to the use of environmental degradation studies. Follow-up ecotoxicological studies should then be conducted using the resulting materials identified in the degradation studies. There is a need to develop certified standard reference materials that are environmentally relevant and meet the needs for risk assessment. This would help to characterize the effects of environmentally relevant plastics. Experiments that consider chronic effects (including effects of long-term retention within organisms) using consistent endpoints should also be completed. Provencher et al. (2018b) highlighted a need for studies that examine plastic transfer between predator and prey, as well as the biomagnification, bioaccumulation, and bioconcentration of these transferred plastics. Further research is needed on the mechanisms of absorption, distribution, metabolism, and excretion of microplastics and on the feasibility of a read-across approach from particle translocation studies. There is also a need to develop a better understanding of the sublethal, interactive and cumulative effects of plastics with other factors. For example, although a recent study has shown that there may be sublethal effects related to plastic ingestion on the blood chemistry of flesh-footed shearwaters in the southern hemisphere, the authors are unable to make definitive links at this time (Lavers et al. 2019). Further, while studies such as those by Lavers and Bond (2016) on ingested plastics as a route for the transport of trace metals have indicated that concentrations of certain metals were positively related to plastic mass, generalizations about the transfer of trace elements from ingested plastics are not yet possible as the mechanisms underlying this process are unknown. In addition, some studies on microplastic ingestion have only examined a portion of an organism's digestive tract, which may lead to an underestimation of ingestion rates, since other components of the GI tract may also contain microplastic particles. To accurately estimate all ingested microplastic, it is recommended that the entire GI tract, from esophagus to vent, of fish and the entire body for smaller species (e.g., bivalves) be examined (Hermsen et al. 2018).

There is a lack of studies on microplastics in soil, and further research is needed to fully understand the interactive effects that plastic pollution will have on soil fauna and potential uptake into food crops. Finally, while some sources and occurrences of microfibres have been identified, further work is needed to fully understand their distribution and fate in the environment, as well as the effects this type of plastic pollutant presents. 

Recent research has begun to explore links between plastic pollution and climate change. For example, Royer et al. (2018) showed that commonly used plastics produce greenhouse gases when exposed to ambient solar radiation, and virgin plastics had higher emissions of hydrocarbon gases than environmentally aged plastic pellets. This suggests that plastic pollution may be contributing to climate change. There is also evidence to suggest that climate change could contribute to increased wildlife exposure to plastic pollution. For example, Drever et al. (2018) reported that, under conditions of unusually warm ocean temperatures, red phalaropes (Phalaropus fulicarius) were found feeding closer to shore. The authors indicated that distribution shifts of the birds closer to shore resulted in increased exposure to plastic pollution.

In addition to the uncertainties inherent in microplastic toxicity testing described above, there is a need to conduct toxicity tests on nanoplastics; however, these may also be confounded by the suspension matrix used (Pikuda et al. 2019). Toxicity results for studies using commercially formulated nanoplastics, which are likely to contain preservatives, antimicrobials, or surfactants, must therefore be carefully considered (Pikuda et al. 2019).

9.3 Human health effects

In order to better understand the potential human health effects of microplastics resulting from both oral and inhalation exposures, an improved understanding of the extent and nature of human exposure and potential toxicological hazards is required. 

With respect to the potential human health impacts of microplastic ingestion (e.g., from drinking water and/or food) and inhalation (e.g., from indoor and ambient air), more research is needed on the uptake and fate of microplastics in the GI and respiratory tracts and on the bioavailability of chemical substances associated with microplastics. In addition, from an inhalation perspective, there is a need to better characterize microplastics exposure for particles of aerodynamic diameter in the micron scale (<1 mm), with a focus on inhalable particles (<10 µm) and especially respirable particles (<2.5 µm) that can penetrate deep into the lungs. There is also a need to understand the physical characteristics of microplastics (e.g., length, diameter, polymer type and surface chemistry) that may determine their bioavailability, tissue distribution, and potential relevance to human health.

Toxicological research using appropriate cell models and experimental animals is needed to better inform human health risk assessment, including identifying target tissues, threshold doses, and mode of action. Epidemiological studies in the general population would also help to inform the human health impacts of microplastics. More research is also required to improve the understanding of whether the characteristics (e.g., size, shape, composition) of microplastics influence their potential adverse effects. In addition, as information on the health-relevant properties of microplastics emerges, standardized reporting metrics are needed to ensure that those features are adequately characterized in scientific reports.

There is also a need to understand the extent to which microplastics may act as a vector for transporting other chemicals (e.g., chemicals additives, adsorbed environmental contaminants) and to determine whether they have an impact on human health. While recent reviews indicate that there is a low health concern for human exposure to chemicals from ingestion of microplastics from food or drinking water (EFSA 2016; FAO 2017; WHO 2019), further research would be required before a human health risk assessment on microplastics is possible. Further research investigating the toxicity of nanoplastics is also required, as described above.

Lastly, there is also a need for improved characterization of microplastic-associated biofilms in drinking water, drinking water sources and air. Gaining increased knowledge in areas such as the factors shaping biofilm composition, the taxonomy of biofilm communities, and biofilm activity and interactions (e.g., transfer of ARGs) would contribute to the understanding of the importance of biofilms on human health.  

10. Findings

Plastic pollution, in the form of macroplastics and microplastics, is ubiquitous in the environment. It is estimated that in 2016, 1% of all plastic waste in Canada, or 29 kt, was discharged to the environment as pollution. Since plastics degrade very slowly and are persistent in the environment, the frequency of occurrence of plastic pollution in the environment is expected to increase. 

Macroplastics have been demonstrated to cause physical harm to environmental receptors on an individual level and to have the potential to adversely affect habitat integrity. Organisms have been shown to ingest macroplastics and to become entangled in macroplastics, which can result in direct harm and in many cases, mortality. 

The evidence for potential effects of microplastic pollution on environmental receptors is less clear and sometimes contradictory, and further research is required. For example, although there are reports indicating that exposure of environmental receptors to microplastics can lead to mortality, developmental and reproductive effects, effects on feeding and energy production, and biochemical or molecular-level effects, a similar number of reports have found no effects. 

The current literature on the human health effects of microplastics is limited, although a concern for human health has not been identified at this time. Potential exposure pathways include air, water and food. While some occupational epidemiology and experimental animal studies show the potential for effects at high exposure concentrations, they are of questionable reliability and relevance, and further research on the potential for microplastics to impact human health is required.

In order to advance the understanding of the impacts of plastic pollution on the environment and human health, it is recommended that research be conducted to address key knowledge gaps identified in this report. This includes studies to improve the understanding of both exposure to and potential toxicity of plastics. More specifically, research is recommended in the following areas:

• Developing standardized methods for sampling, quantifying, characterizing and evaluating the effects of macroplastics and microplastics;
• Furthering understanding of human exposure to microplastics;
• Furthering understanding of the ecotoxicological effects of microplastics;
• Furthering understanding of the effects of microplastics on human health; and
• Expanding and developing consistent monitoring efforts to include lesser characterized environmental compartments such as soil.

Given the increasing amounts of plastic pollution in the environment and the demonstrated ability of macroplastics to harm biota, it is anticipated that the frequency of occurrence of physical effects on individual environmental receptors will continue to increase if current trends continue without mitigation measures. 

In accordance with the precautionary principle, action is needed to reduce macroplastics and microplastics that end up in the environment.

11. References

Abbasi S, Soltani N, Keshavarzi B, Moore F, Turner A, Hassanaghaei M. 2018. Microplastics in different tissues of fish and prawn from the Musa Estuary, Persian Gulf. Chemosphere. 205:80-87.

Adam V, Yang T, Nowack B. 2019. Toward an ecotoxicological risk assessment of microplastics: Comparison of available hazard and exposure data in freshwaters. Environ Toxicol Chem. 38(2):436-447. 

AET Group (Kitchener, ON). 2016. 2016 Toronto Litter Audit. Summary Report. Toronto (ON): City of Toronto. 90 p.

AET Group (Kitchener, ON). 2019. 2019 City of Edmonton Litter Audit Report. Final Report. Edmonton (AB): City of Edmonton. 72 p.

Agarwal DK, Kaw JL, Srivastava SP, Seth PK. 1978. Some biochemical and histopathological changes induced by polyvinyl chloride dust in rat lung. Environ Res. 16(1-3):333-341.

Akhbarizadeh R, Moore F, Keshavarzi B. 2018. Investigating a probable relationship between microplastics and potentially toxic elements in fish muscles from northeast of Persian Gulf. Environ Pollut. 232:154-163. 

Akhbarizadeh R, Moore F, Keshavarzi B. 2019. Investigating microplastics bioaccumulation and biomagnification in seafood from the Persian Gulf: a threat to human health? Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 36(11):1696-1708.

Alimi OS, Farner Budarz J, Hernandez LM, Tufenkji N. 2018. Microplastics and nanoplastics in aquatic environments: Aggregation, deposition, and enhanced contaminant transport. Environ Sci Technol. 52(4):1704-1724. 

Aljaibachi R, Callaghan A. 2018. Impact of polystyrene microplastics on Daphnia magna mortality and reproduction in relation to food availability. PeerJ. 6:e4601. 

Allen S, Allen D, Phoenix VR, Le Roux G, Durántez Jiménez P, Simonneau A, Binet S, Galop D. 2019. Atmospheric transport and deposition of microplastics in a remote mountain catchment. Nat Geosci. 12:339-344. 

Amalfitano S, Coci M, Corno G, Luna, GM. 2014. A microbial perspective on biological invasions in aquatic ecosystems. Hydrobiologia. 746(1):13-22. 

[AMAP] Arctic Monitoring and Assessment Programme. 2004. AMAP Assessment 2002: Persistent Organic Pollutants in the Arctic. Oslo (NO): AMAP. 310 p.

Anderson JC, Park BJ, Palace VP. 2016. Microplastics in aquatic environments: implications for Canadian ecosystems. Environ Pollut. 218:269-280.

Anderson PJ, Warrack S, Langen V, Challis JK, Hanson ML, Rennie MD. 2017. Microplastic contamination in Lake Winnipeg, Canada. Environ Pollut. 225:223-231.

Andrady AL. 2015. Persistence of plastic litter in the oceans. In: Bergmann M, Gutow L, Klages M, editors. Marine anthropogenic litter. Cham (CH): Springer. Chapter 3. p. 57-72.

Andrady AL. 2017. The plastic in microplastics: A review. Mar Pollut Bull. 119(1):12-22.

Arias-Andres M, Kettner MT, Miki T, Grossart HP. 2018. Microplastics: New substrates for heterotrophic activity contribute to altering organic matter cycles in aquatic ecosystems. Sci Total Environ. 635:1152-1159.

Arias-Andres M, Rojas-Jimenez K, Grossart H. 2019. Collateral effects of microplastic pollution on aquatic microorganisms: An ecological perspective. Trends Analyt Chem. 112:234-240. 

Araújo PHH, Sayer C, Giudici R, Poço JGR. 2002. Technique for reducing residual monomer content in polymers: a review. Polym Eng Sci. 42(7):1442-1468. [cited in WHO 2019].

Ašmonaitė G, Sundh H, Asker N, Carney Almroth B. 2018. Rainbow trout maintain intestinal transport and barrier functions following exposure to polystyrene microplastics. Environ Sci Technol. 52(24):14392-14401.

Au SY, Bruce TF, Bridges WC, Klaine SJ. 2015. Responses of Hyalella azteca to acute and chronic microplastic exposures. Environ Toxicol Chem. 34(11):2564-2572.

Avio CG, Gorbi S, Milan M, Benedetti M, Fattorini D, d'Errico G, Pauletto M, Bargelloni L, Regoli F. 2015. Pollutants bioavailability and toxicological risk from microplastics to marine mussels. Environ Pollut. 198:211-222.

Baker-Austin C, Wright MS, Stepanauskas R, McArthur JV. 2006. Co-selection of antibiotic and metal resistance. Trends Microbiol. 14(4):176-182.

Bakir A, O'Connor IA, Rowland SJ, Hendriks AJ, Thompson RC. 2016. Relative importance of microplastics as a pathway for the transfer of hydrophobic organic chemicals to marine life. Environ Pollut. 219:56-65.

Baldwin AK, Corsi SR, Mason SA. 2016. Plastic debris in 29 Great Lakes tributaries: Relations to watershed attributes and hydrology. Environ Sci Technol. 50(19):10377-10385.

Balestri E, Menicagli V, Ligorini V, Fulignati S, Raspolli Galletti AM, Lardicci C. 2019. Phytotoxicity assessment of conventional and biodegradable plastic bags using seed germination test. Ecol Indicators. 102:569-580.

Ball H, Cross R, Grove E, Horton A, Johnson A, Jürgens M, Read D, Svendsen C. 2019. Sink to river - river to tap: A review of potential risks from nanoparticles and microplastics. London (GB): UK Water Industry Research Limited. Report No.:19/EQ/01/18.

Ballent A, Corcoran PL, Madden O, Helm PA, Longstaffe FJ. 2016. Sources and sinks of microplastics in Canadian Lake Ontario nearshore, tributary and beach sediments. Mar Pollut Bull. 110(1):383-395.

Barboza LGA, Dick Vethaak A, Lavorante BRBO, Lundebye AK, Guilhermino L. 2018. Marine microplastic debris: An emerging issue for food security, food safety and human health. Mar Pollut Bull. 133:336-348.

Barboza LGA, Cunha SC, Monteiro C, Fernandes JO, Guilhermino L. 2020. Bisphenol A and its analogs in muscle and liver of fish from the North East Atlantic Ocean in relation to microplastic contamination. Exposure and risk to human consumption. J Hazard Mater. 393:122419.

Batel A, Linti F, Scherer M, Erdinger L, Braunbeck T. 2016. Transfer of benzo[a]pyrene from microplastics to Artemia nauplii and further to zebrafish via a trophic food web experiment: CYP1A induction and visual tracking of persistent organic pollutants. Environ Toxicol Chem. 35(7):1656-1666.

Batté M, Appenzeller BMR, Grandjean D, Fass S, Gauthier V, Jorand F, Mathieu L, Boualam M, Saby S, Block JC. 2003. Biofilms in drinking water distribution systems. Rev Environ Sci Biotechnol. 2(2):147-168.

Beer S, Garm A, Huwer B, Dierking J, Nielsen TG. 2018. No increase in marine microplastic concentration over the last three decades—A case study from the Baltic Sea. Sci Total Environ. 621:1272-1279.

Beiras R, Bellas J, Cachot J, Cormier B, Cousin X, Engwall M, Gambardella C, Garaventa F, Keiter S, Le Bihanic F, et al. 2018. Ingestion and contact with polyethylene microplastics does not cause acute toxicity on marine zooplankton. J Hazard Mater. 360:452-460. 

Bergmann M, Wirzberger B, Krumpen T, Lorenz C, Primpke S, Tekman MB, Gerdts G. 2017. High quantities of microplastic in Arctic deep-sea sediments from the HAUSGARTEN Observatory. Environ Sci Technol. 51:11000-11010. 

Bernardini I, Garibaldi F, Canesi L, Fossi MC, Baini M. 2018. First data on plastic ingestion by blue sharks (Prionace glauca) from the Ligurian Sea (North-Western Mediterranean Sea). Mar Pollut Bull. 135:303-310. 

Berry D, Xi C, Raskin L. 2006. Microbial ecology of drinking water distribution systems. Curr Opin Biotechnol. 17(3):297-302.

Besseling E, Redondo-Hasselerharm P, Foekema EM, Koelmans AA. 2019. Quantifying ecological risks of aquatic micro- and nanoplastic. Crit Rev Environ Sci Technol. 49:32-80.

Bhardwaj H, Gupta R, Tiwari A. 2013. Communities of microbial enzymes associated with biodegradation of plastics. J Polym Environ. 21(2):575-579.

Biber NFA, Foggo A, Thompson RC. 2019. Characterising the deterioration of different plastics in air and seawater. Mar Pollut Bull. 141:595-602.

Bishop DP. 1995. Physical and chemical effects of domestic laundering processes. In: Carr CM, editor. Chemistry of the Textiles Industry. Glasgow (GB): Springer, Dordrecht. p. 125-172.

Blašković A, Fastelli P, Čižmek H, Guerranti C, Renzi M. 2017. Plastic litter in sediments from the Croatian marine protected area of the natural park of Telaščica bay (Adriatic Sea). Mar Pollut Bull. 114(1):583-586.

Blarer P, Burkhardt-Holm P. 2016. Microplastics affect assimilation efficiency in the freshwater amphipod Gammarus fossarum. Environ Sci Pollut Res. 23(23):23522-23532.

Bläsing M, Amelung W. 2018. Plastics in soil: Analytical methods and possible sources. Sci Total Environ. 612:422-435.

Blettler MCM, Ulla MA, Rabuffetti AP, Garello N. 2017. Plastic pollution in freshwater ecosystems: macro‑, meso-, and microplastic debris in a floodplain lake. Environ Monit Assess. 189(11):581.

Böhmert L, Stock V, Braeuning A. 2019. Plausibility of microplastic uptake in a paper by Deng et al., Scientific Reports 7: 46687, 2017. Arch Toxicol. 93(1):217-218.

Boots B, Russell CW, Green DS. 2019. Effects of microplastics in soil ecosystems: above and below ground. Environ Sci Technol. 53:11496-11506.

Bordós G, Urbányi B, Micsinai A, Kriszt B, Polatai Z, Szabó I, Hantosi Z, Szoboszlay S. 2019. Identification of microplastics in fish ponds and natural freshwater environments of the Carpathian basin, Europe. Chemosphere. 216:110-116.

Borm PA, Driscoll KE. 2019. The hazards and risks of inhaled poorly soluble particles - where do we stand after 30 years of research? Part Fibre Toxicol. 16(1):11.

Boucher J, Friot D. 2017. Primary microplastics in the oceans: A global evaluation of sources. Gland, (CH): [IUCN] International Union for Conservation of Nature. 43 p. 

Bour A, Haarr A, Keiter S, Hylland K. 2018. Environmentally relevant microplastic exposure affects sediment-dwelling bivalves. Environ Pollut. 236:652-660. 

Bourdages MPT, Provencher JF, Sudlovenick E, Ferguson SH, Young BG, Pelletier N, Murphy MJJ, D'Addario A, Vermaire JC. 2020. No plastics detected in seal (Phocidae) stomachs harvested in the eastern Canadian Arctic. Mar Pollut Bull. 150:110772.

Bouwman H, Minnaar K, Bezuidenhout C, Verster C. 2018. Microplastics in freshwater environments – A Scoping Study. Technical Report. Republic of South Africa: Water Research Commission. 62 p. Report No. 2610/1/18.

Braeuning A. 2019. Uptake of microplastics and related health effects: a critical discussion of Deng et al., Scientific reports 7:46687, 2017. Arch Toxicol. 93(1):219-220.

Brandão ML, Braga KM, Luque JL. 2011. Marine debris ingestion by Magellenic penguins, Spheniscus magellanicus, from the Brazilian coastal zone. Mar Pollut Bull. 62(10):2246-2249. 

Brennecke D, Duarte B, Paiva F, Caçador I, Canning-Clode J. 2016. Microplastics as vector for heavy metal contamination from the marine environment. Estuar Coast Shelf Sci. 178:189-195.

Brodie EL, DeSantis TZ, Parker JP, Zubietta IX, Piceno YM, Andersen GL. 2007. Urban aerosols harbor diverse and dynamic bacterial populations. Proc Natl Acad Sci U S A. 104(1):299-304. [cited in Prata 2018].

Brown DM, Wilson MR, MacNee W, Stone V, Donaldson K. 2001. Size-dependent proinflammatory effects of ultrafine polystyrene particles: a role for surface area and oxidative stress in the enhanced activity of ultrafines. Toxicol Appl Pharmacol. 175(3):191-199.

Browne MA, Dissanayake A, Galloway TS, Lowe DM, Thompson RC. 2008. Ingested microscopic plastic translocates to the circulatory system of the mussel, Mytilus edulis (L.). Environ Sci Technol. 42(13):5026-5031.

Browne MA, Galloway TS, Thompson RC. 2010. Spatial Patterns of Plastic Debris along Estuarine Shorelines. Environ Sci Technol. 44(9):3404-3409.

Burns EE, Boxall ABA. 2018. Microplastics in the aquatic environment: Evidence for or against adverse impacts and major knowledge gaps. Environ Toxicol Chem. 37(11):2776-2796. 

Cai L, Wang J, Peng J, Tan Z, Zhan Z, Tan X, Chen Q. 2017. Characteristic of microplastics in the atmospheric fallout from Dongguan City, China: Preliminary research and first evidence. Environ Sci Pollut Res Int. 24(32):24928-24935. 

Campbell SH, Williamson PR, Hall BD. 2017. Microplastics in the gastrointestinal tracts of fish and the water from an urban prairie creek. Facets (Ott). 2:395-409. 

Carney Almroth BM, Åström L, Roslund S, Petersson H, Johansson M, Persson NK. 2018. Quantifying shedding of synthetic fibres from textiles; a source of microplastics released into the environment. Environ Sci Pollut Res Int. 25(2):1191-1199. 

Catarino AI, Macchia V, Sanderson WG, Thompson RC, Henry TB. 2018. Low levels of microplastics (MP) in wild mussels indicate that MP ingestion by humans is minimal compared to exposure via household fibres fallout during a meal. Environ Pollut. 237:675-684.

Cau A, Avio CG, Dessì C, Follesa MC, Moccia D, Regoli F, Pusceddu A. 2019. Microplastics in the crustaceans Nephrops norvegicus and Aristeus antennatus: Flagship species for deep-sea environments? Environ Pollut. 255(Pt 1):113107.

[CBD] Secretariat of the Convention on Biological Diversity. 2012. Impacts of Marine Debris on Biodiversity: Current Status and Potential Solutions. Technical Series No. 67. Montreal (QC): CBD. 61 p. 

[CBD] Secretariat of the Convention on Biological Diversity. 2016. Marine debris: Understanding, preventing, and mitigating the significant adverse impacts on marine and coastal biodiversity. Technical Series No. 83. Montreal (QC): CBD. 78 p. 

[CCME] Canadian Council of Ministers of the Environment. 2018. Strategy on Zero Plastic Waste. Winnipeg (MB): CCME. 14 p. Report No. PN 1583.

Cesa FS, Turra A, Checon HH, Leonardi B, Baruque-Ramos J. 2020. Laundering and textile parameters influence fibers release in household washings. Environ Pollut. 257:113553.

Chae Y, Hong SH, An Y-J. 2020. Photosynthesis enhancement in four marine microalgal species exposed to expanded polystyrene leachate. Ecotoxicol Environ Saf. 189:109936.

Chagnon C, Thiel M, Antunes J, Ferreira JL, Sobral P, Ory NC. 2018. Plastic ingestion and trophic transfer between Easter Island flying fish (Cheilopogon rapanouiensis) and yellowfin tuna (Thunnus albacares) from Rapa Nui (Easter Island). Environ Pollut. 243(Pt A):127-133.

Chapron L, Peru E, Engler A, Ghiglione JF, Meistertzheim AL, Pruski AM, Purser A, Vétion G, Galand PE, Lartaud F. 2018. Macro- and microplastics affect cold-water corals growth, feeding, and behaviour. Sci Rep. 8(1):15299. 

Chen H, Wang S, Guo H, Lin H, Zhang Y, Long Z, Huang H. 2019. Study of marine debris around a tourist city in East China: Implication for waste management. Sci Total Environ. 676:278-289.

Chen H, Wang Y, Sun X, Peng Y, Xiao L. 2020. Mixing effect of polylactic acid microplastic and straw residue on soil property and ecological function. Chemosphere. 243:125271.

Chen J, Tan M, Nemmar A, Song W, Dong M, Zhang G, Li Y. 2006. Quantification of extrapulmonary translocation of intratracheal-instilled particles in vivo in rats: effect of lipopolysaccharide. Toxicology. 222(3):195-201.

Chen Q, Allgeier A, Yin D, Hollert H. 2019. Leaching of endocrine disrupting chemicals from marine microplastics and mesoplastics under common life stress conditions. Environ Int. 130:104938.

Chen Y, Liu X, Leng Y, Wang J. 2020. Defense responses in earthworms (Eisenia fetida) exposed to low-density polyethylene microplastics in soils. Ecotoxicol Environ Saf. 187:109788.

Choy CA, Drazen JC. 2013. Plastic for dinner? Observations of frequent debris ingestion by pelagic predatory fishes from the central North Pacific. Mar Ecol Prog Ser. 485:155-163. 

Chiba S, Saito H, Fletcher R, Yogi T, Kayo M, Miyagi S, Ogido M, Fujikura K. 2018. Human footprint in the abyss: 30 year records of deep-sea plastic debris. Mar Policy. 96:204-212.

Chua EM, Shimeta J, Nugegoda D, Morrison PD, Clarke BO. 2014. Assimilation of polybrominated diphenyl ethers from microplastics by the marine amphipod, Allorchestes compressa. Environ Sci Technol. 48(14):8127-8134.

Cole M, Lindeque PK, Fileman E, Halsband C, Galloway TS. 2015. The impact of polystyrene microplastics on feeding, function and fecundity in the marine copepod Calaus helgolandicus. Environ Sci Technol. 49(2):1130-1137. 

Cole M, Lindeque PK, Fileman E, Clark J, Lewis C, Halsband C, Galloway TS. 2016. Microplastics alter the properties and sinking rates of zooplankton faecal pellets. Environ Sci Technol. 50:3239-3246.

Collicutt B, Juanes F, Dudas SE. 2019. Microplastics in juvenile Chinook salmon and their nearshore environments on the east coast of Vancouver Island. Environ Pollut. 244:135-142. 

Cong Y, Jin F, Tian M, Wang J, Shi H, Wang Y, Mu J. 2019. Ingestion, egestion and post-exposure effects of polystyrene microspheres on marine medaka (Oryzias melastigma). Chemosphere. 228:93-100.

Connors EJ. 2017. Distribution and biological implications of plastic pollution on the fringing reef of Mo'orea, French Polynesia. PeerJ. 5:e3733.

Cook L, Chatterjee A, Barnes A, Yarwood J, Hu WS, Dunny G. 2011. Biofilm growth alters regulation of conjugation by a bacterial pheromone. Mol Microbiol. 81(6):1499-1510.

Corcoran PL, Norris T, Ceccanese T, Walzak MJ, Helm PA, Marvin CH. 2015. Hidden plastics of Lake Ontario, Canada and their potential preservation in the sediment record. Environ Pollut. 204:17-25.

Corcoran PL, Belontz SL, Ryan K, Walzak MJ. 2020. Factors controlling the distribution of microplastic particles in benthic sediment of the Thames River, Canada. Environ Sci Technol. 54:818-825.

Corradini F, Meza P, Eguiluz R, Casado F, Huerta-Lwanga E, Geissen V. 2019. Evidence of microplastic accumulation in agricultural soils from sewage sludge disposal. Sci Total Environ. 671:411-420.

Cosgrove L, McGeechan PL, Robson GD, Handley PS. 2007. Fungal communities associated with degradation of polyester polyurethane in soil. Appl Environ Microbiol. 73(18):5817-8524.

Covernton GA, Pearce CM, Gurney-Smith HJ, Chastain SG, Ross PS, Dower JF, Dudas SE. 2019. Size and shape matter: A preliminary analysis of microplastic sampling technique in seawater studies with implications for ecological risk assessment. Sci Total Environ. 667:124-132.

Cox KD, Covernton GA, Davies HL, Dower JF, Juanes F, Dudas SE. 2019. Human consumption of microplastics. Environ Sci Technol. 53(12):7068-7074.

Crittenden JC, Rhodes Trussell R, Hand DW, Howe KJ, Tchobanoglous G. 2012. MWH's Water Treatment Principles and Design. Third Edition. Hoboken (NJ): Wiley & Sons Inc. 1920 p.

Crossman J, Hurley RR, Futter M, Nizzetto L. 2020. Transfer and transport of microplastics from biosolids to agricultural soils and the wider environment. Sci Total Environ. 724:138334.

Curren E, Leong SCY. 2019. Profiles of bacterial assemblages from microplastics of tropical coastal environments. Sci Total Environ. 655:313-320.

Cuthbert RN, Al-Jaibachi R, Dalu T, Dick JTA, Callaghan A. 2019. The influence of microplastics on trophic interaction strengths and oviposition preferences of dipterans. Sci Total Environ. 651(Pt 2):2420-2423. 

da Costa JP, Paço A, Santos PSM, Duarte AC, Rocha-Santos T. 2019. Microplastics in soils: assessment, analytics and risks. Environ Chem. 16(1):18-30.

da Costa Araújo AP, Gomes A, Malafaia G. 2020. Hepatotoxicity of pristine polyethylene microplastics in neotropical physalaemus cuvieri tadpoles (Fitzinger, 1826). J Hazard Mater. 386:121992.

Dai Z, Zhang H, Zhou Q, Tian Y, Chen T, Tu C, Fu C, Luo Y. 2018. Occurrence of microplastics in the water column and sediment in an inland sea affected by intensive anthropogenic activities. Environ Poll. 242(Pt B):1557-1565.

Dawson AL, Kawaguchi S, King CK, Townsend KA, King R, Huston WM, Bengtson Nash SM. 2018. Turning microplastics into nanoplastics through digestive fragmentation by Antarctic krill. Nat Commun. 9:1001.

de Haan WP, Sanchez-Vidal A, Canals M, NUREIEV1 Shipboard Scientific Party. 2019. Floating microplastics and aggregate formation in the Western Mediterranean Sea. Mar Pollut Bull. 140:523-535.

De Falco F, Gullo MP, Gentile G, Di Pace E, Cocca M, Gelabert L, Brouta-Agnésa M, Rovira A, Escudero R, Villalba R, et al. 2018. Evaluation of microplastic release caused by textile washing processes of synthetic fabrics. Environ Pollut. 236:916-925. 

De Falco F, Cocca M, Avella M, Thompson RC. 2020. Microfiber release to water, via laundering, and to air, via everyday use: A comparison between polyester clothing with differing textile parameters. Environ Sci Technol. 54(6):3288-3296.

De Felice B, Bacchetta R, Santo N, Tremolada P, Parolini M. 2018. Polystyrene microplastics did not affect body growth and swimming activity in Xenopus laevis tadpoles. Environ Sci Pollut Res Int. 25(34):34644-34651.

De Felice B, Sabatini V, Antenucci S, Gattoni G, Santo N, Bacchetta R, Ortenzi MA, Parolini M. 2019. Polystyrene microplastics ingestion induced behavioral effects to the cladoceran Daphnia magna. Chemosphere. 231:423-431.

de Scisciolo T, Mijts EN, Becker T, Eppinga MB. 2016. Beach debris on Aruba, Southern Caribbean: Attribution to local land-based and distal marine-based sources. Mar Pollut Bull. 106(1-2):49-57.

de Souza Machado AA, Lau CW, Kloas W, Bergmann J, Bachelier JB, Faltin E, Becker R, Görlich AS, Rillig MC. 2019. Microplastics can change soil properties and affect plant performance. Environ Sci Technol. 53(10):6044-6052.

De Tender CA, Devriese LI, Haegeman A, Maes S, Ruttink T, Dawyndt P. 2015. Bacterial community profiling of plastic litter in the Belgian part of the North Sea. Environ Sci Technol. 49(16):9629-9638.

de Vries AN, Govoni D, Árnason SH, Carlsson P. 2020. Microplastic ingestion by fish: Body size, condition factor and gut fullness are not related to the amount of plastics consumed. Mar Pollut Bull. 151:110827. 

De Witte B, Devriese L, Bekaert K, Hoffman S, Vandermeersch G, Cooreman K, Robbens J. 2014. Quality assessment of the blue mussel (Mytilus edulis): comparison between commercial and wild types. Mar Pollut Bull. 85(1):146-155.

Dean BY, Corcoran PL, Helm PA. 2018. Factors influencing microplastic abundances in nearshore, tributary and beach sediments along the Ontario shoreline of Lake Erie. J Great Lakes Res. 44(5):1002-1009.

Delacuvellerie A, Cyriaque V, Gobert S, Benali S, Wattiez R. 2019. The plastisphere in marine ecosystem hosts potential specific microbial degraders including Alcanivorax borkumensis as a key player for the low-density polyethylene degradation. J Hazard Mater. 380:120899.

DeMatteo R, Keith MM, Brophy JT, Wordsworth A, Watterson AE, Beck M, Ford AR, Gilbertson M, Pharityal J, Rootham M, et al. 2012. Chemical exposures of women workers in the plastics industry with particular reference to breast cancer and reproductive hazards. New Solut. 22(4):427-448.

Demers PA, Schade WJ, Demers RY. 1994. Lymphocytopenia and occupational exposures among pattern and model makers. Scand J Work Environ Health. 20(2):107-112.

Deng Y, Zhang Y, Lemos B, Ren H. 2017. Tissue accumulation of microplastics in mice and biomarker responses suggest widespread health risks of exposure. Sci Rep. 7:46687.

Deng Y, Zhang Y, Qiao R, Bonilla MM, Yang X, Ren H, Lemos B. 2018. Evidence that microplastics aggravate the toxicity of organophosphorus flame retardants in mice (Mus musculus). J Hazard Mater. 357:348-354.

Desforges J-PW, Galbraith M, Ross PS. 2015. Ingestion of microplastics by zooplankton in the northeast Pacific Ocean. Archi Environ Contam Toxicol. 69(3):320-330.

Devriese LI, van der Meulen MD, Maes T, Bekaert K, Paul-Pont I, Frère L, Robbens J, Vethaak AD. 2015. Microplastic contamination in brown shrimp (Crangon crangon, Linnaeus 1758) from coastal waters of the Southern North Sea and Channel area. Mar Pollut Bull. 98(1-2):179-187.

Devriese LI, De Witte B, Vethaak AD, Hostens K, Leslie HA. 2017. Bioaccumulation of PCBs from microplastics in Norway lobster (Nephrops norvegicus): An experimental study. Chemosphere. 186:10-16. 

Di M, Wang J. 2018. Microplastics in surface waters and sediments of the three gorges reservoir, China. Sci Total Environ. 616-617:1620-1627.

Diepens NJ, Koelmans AA. 2018. Accumulation of plastic debris and associated contaminants in aquatic food webs. Environ Sci Technol. 52(15):8510-8520. 

Dillon Consulting. 2019. Street Litter Audits – 2019 Results. Final Report. Vancouver (BC): City of Vancouver. 71 p.

Donaldson K. 2009. The inhalation toxicology of p-aramid fibrils. Crit Rev Toxicol. 39(6):487-500.

Doyle D, Frias J, Nash R, Gammell M. 2020. Current environmental microplastic levels do not alter emergence behaviour in the intertidal gastropod Littorina littorea. Mar Pollut Bull. 151:110859.

Drever MC, Provencher JF, O’Hara PD, Wilson L, Bowes V, Bergman CM. 2018. Are ocean conditions and plastic debris resulting in a “double whammy” for marine birds? Mar Pollut Bull. 133:684-692.

Driedger AGJ, Dürr HH, Mitchell K, Van Cappellen P. 2015. Plastic debris in the Laurentian Great Lakes: A review. J Great Lakes Res. 41(1):9-19.

Dris R, Gasperi J, Rocher V, Saad M, Renault N, Tassin B. 2015. Microplastic contamination in an urban area: A case study in Greater Paris. Environ Chem. 12(5):592-599. 

Dris R, Gasperi J, Saad M, Mirande C, Tassin B. 2016. Synthetic fibers in atmospheric fallout: A source of microplastics in the environment? Mar Pollut Bull. 104(1-2):290-293.

Dris R, Gasperi J, Mirande C, Mandin C, Guerrouache M, Langlois V, Tassin B. 2017. A first overview of textile fibers, including microplastics, in indoor and outdoor environments. Environ Pollut. 221:453-458.

Duncan EM, Arrowsmith JA, Bain CE, Bowdery H, Broderick AC, Chalmers T, Fuller WJ, Galloway TS, Lee JH, Lindeque PK, et al. 2019. Diet-related selectivity of macroplastic ingestion in green turtles (Chelonia mydas) in the eastern Mediterranean. Sci Rep. 9(1):11581.

Earthwatch Europe. 2018. Plastic Rivers reducing the plastic pollution on our doorstep [PDF]. Oxford (GB): Earthwatch Institute. 20 p. [accessed 2019 May 15].

[ECCC] Environment and Climate Change Canada. 2015. Microbeads – A Science Summary. Ottawa (ON): Government of Canada. 34 p.

[ECCC] Environment and Climate Change Canada. 2019b. Canada’s Plastics Science Agenda. Ottawa (ON): Government of Canada. 23 p.

 [ECCC] Environment and Climate Change Canada. 2019a. Economic study of the Canadian plastic industry, markets and waste. Summary Report. Ottawa (ON): Government of Canada. 43 p.

[ECCC] Environment and Climate Change Canada. 2019c. Economic study of the Canadian plastic industry, markets and waste. Final Report - Task 1. Ottawa (ON): Government of Canada. 186 p. Available from: ec.pddemandes-pd-requests.ec@canada.ca.

 [ECCC] Environment and Climate Change Canada. 2019d. Economic study of the Canadian plastic industry, markets and waste. Final Report - Task 2. Ottawa (ON): Government of Canada. 121 p. Available from: ec.pddemandes-pd-requests.ec@canada.ca.

[ECCC] Environment and Climate Change Canada. 2019e. Economic study of the Canadian plastic industry, markets and waste. Final Report - Task 5. Ottawa (ON): Government of Canada. 62 p. Available from: ec.pddemandes-pd-requests.ec@canada.ca.

[ECHA] European Chemicals Agency. 2019. ANNEX to the ANNEX XV Restriction Report - Proposal for a restriction on intentionally added microplastics. Helsinki (FI): ECHA. 299 p.

Eckert EM, Di Cesare A, Kettner MT, Arias-Andres M, Fontaneto D, Grossart HP, Corno G. 2018a. Microplastics increase impact of treated wastewater on freshwater microbial community. Environ Pollut. 234:495-502.

Eckert EM, Di Cesare, A, Coci M, Corno G. 2018b. Persistence of antibiotic resistance genes in large subalpine lakes: The role of anthropogenic pollution and ecological interactions. Hydrobiologia. 824(1):93-108. 

[EFSA] European Food Safety Authority. 2016. Presence of microplastics and nanoplastics in food, with particular focus on seafood. EFSA J. 14(6):e04501.

Egeland GM and [CINE] Centre for Indigenous Peoples’ Nutrition and Environment. 2010a. Inuit Health Survey 2007-2008 - Inuvialuit Settlement Region. [PDF] Ste-Anne-de-Bellevue (QC): CINE. 31 p.

Egeland GM and [CINE] Centre for Indigenous Peoples’ Nutrition and Environment. 2010b. Inuit Health Survey 2007-2008 - Nunatsiavut. [PDF] Ste-Anne-de-Bellevue (QC): CINE. 35 p.

Egeland GM and [CINE] Centre for Indigenous Peoples’ Nutrition and Environment. 2010c. Inuit Health Survey 2007-2008 - Nunavut. [PDF] Ste-Anne-de-Bellevue (QC): CINE. 51 p.

Ekebafe LO, Ogbeifun DE, Okieimen FE. 2011. Polymer applications in agriculture. Biokemistri. 23(2):81-89.

Engler RE. 2012. The complex interaction between marine debris and toxic chemicals in the ocean. Environ Sci Technol. 46(22):12302-12315. [cited in Galloway 2015].

Eriksen M, Lebreton LCM, Carson HS, Thiel M, Moore CJ, Borerro JC, Galgani F, Ryan PG, Reisser J. 2014. Plastic pollution in the world’s oceans: More than 5 trillion plastic pieces weighing over 250,000 tons afloat at sea. PLoS One. 9(12):e111913.

Erni-Cassola G, Gibson MI, Thompson RC, Christie-Oleza JA. 2017. Lost, but found with Nile Red: A novel method for detecting and quantifying small microplastics (1 mm to 20 μm) in environmental samples. Environ Sci Technol. 51(23):13641-13648. 

Esiukova E. 2017. Plastic pollution on the Baltic beaches of Kaliningrad region, Russia. Mar Pollut Bull. 114(2):1072-1080.

Espinosa C, Esteban MÁ, Cuesta A. 2019. Dietary administration of PVC and PE microplastics produces histological damage, oxidative stress and immunoregulation in European sea bass. Fish Shellfish Immunol. 95:574-583.

Eubeler JP, Bernhard M, Knepper TP. 2010. Environmental biodegradation of synthetic polymers II. Biodegradation of different polymer groups. Trends Anal Chem. 29:84-100.

European Commission. 2018. Report from the Commission to the European Parliament and the Council on the impact of the use of oxo-degradable plastic, including oxo-degradable plastic carrier bags, on the environment. Final Report. Brussels (BE): European Commission. 8 p.

European Commission. 2019. Environmental and Health Risks of Microplastic Pollution. Independent Expert Report. Brussels (BE): European Commission. 60 p.

Farrell P, Nelson K. 2013. Trophic level transfer of microplastic: Mytilus edulis (L.) to Carcinus maenas (L.). Environ Pollut. 177:1-3.

[FAO] Food and Agriculture Organization of the United Nations. 2017. Microplastics in fisheries and aquaculture: Status of knowledge on their occurrence and implications for aquatic organisms and food safety. Technical Report. Rome (IT): FAO. 126 p. Report No. 615.

Fastelli P, Blašković A, Benardi G, Romeo T, Čižmek H, Andaloro F, Russo GF, Guerranti C, Renzi M. 2016. Plastic litter in sediments from a marine area likely to become protected (Aeolian Archipelago's islands, Tyrrhenian sea). Mar Pollut Bull. 113(1-2):526-529.

Feng Z, Zhang T, Li Y, He X, Wang R, Xu J, Gao G. 2019. The accumulation of microplastics in fish from an important fish farm and mariculture area, Haizhou Bay, China. Sci Total Environ. 696:133948.

Fischer M, Goßmann I, Scholz-Böttcher M. 2019. Fleur de Sel - An interregional monitor for microplastics mass load and composition in European coastal waters? J Anal Appl Pyrolysis. 144:104711. 

Flemming HC, Wingender J. 2010. The biofilm matrix. Nat Rev Microbiol. 8(9):623-633.

Foley CJ, Feiner ZS, Malinich TD, Höök TO. 2018. A meta-analysis of the effects of exposure to microplastics on fish and aquatic invertebrates. Sci Total Environ. 631-632:550-559. 

Fournier S, D’Errico J, Stapleton P. 2018. Translocation of engineered nanomaterials from the maternal lungs to the fetal compartment after instillation. FASEB J. 32(Suppl. 1).  Abstract No. 830.5. 

Frydkjær CK, Iversen N, Roslev P. 2017. Ingestion and egestion of microplastics by the cladoceran Daphnia magna: Effects of regular and irregular shaped plastic and sorbed phenanthrene. Bull Environ Contam Toxicol. 99:655-661.

Gago J, Portela S, Filgueiras AV, Pauly Salinas M, Macías D. 2020. Ingestion of plastic debris (macro and micro) by longnose lancetfish (Alepisaurus ferox) in the North Atlantic Ocean. Reg Stud Mar Sci. 33:100977.

Gall SC, Thompson RC. 2015. The impact of debris on marine life. Mar Pollut Bull. 92(1-2):170-179.

Galloway TS. 2015. Micro- and nano-plastics and human health. In: Bergmann M, Gutow L, Klages M, editors. Marine Anthropogenic Litter. Cham (CH): Springer. Chapter 13. p. 343-366.

Gambardella C, Morgana S, Bramini M, Rotini A, Manfra L, Migliore L, Piazza V, Garaventa F, Faimali M. 2018. Ecotoxicological effects of polystyrene microbeads in a battery of marine organisms belonging to different trophic levels. Mar Environ Res. 141:313-321.

Garcia-Garin O, Vighi M, Sala B, Aguilar A, Tsangaris C, Digka N, Kaberi H, Eljarrat E, Borrell A. 2020. Assessment of organophosphate flame retardants in Mediterranean Boops boops and their relationship to anthropization levels and microplastic ingestion. Chemosphere. 252:126569. 

García-Rivera S, Lizaso JLS, Millán JMB. 2018. Spatial and temporal trends of marine litter in the Spanish Mediterranean seafloor. Mar Pollut Bull. 2018. 137:252-261.

Garrido S, Linares M, Campillo JA, Albentosa M. 2019. Effect of microplastics on the toxicity of chlorpyrifos to the microalgae Isochrysis galbana, clone t-ISO. Ecotoxicol Environ Saf. 173:103-109.

Gasperi J, Dris R, Mirande C, Mandin C, Langlois V, Tassin B. 2015. First overview of microplastics in indoor and outdoor air. In: Paper Presented at 15th EuCheMS International Conference on Chemistry and the Environment (ICCE); 2015 Sep 20-24; Leipzig (DE): ICCE.

Gasperi J, Wright SL, Dris R, Collard F, Mandin C, Guerrouache M, Langlois V, Kelly FJ, Tassin B. 2018. Microplastics in air: Are we breathing it in? Curr Opin Environ Sci Health. 1:1-5.

[GCSC] Great Canadian Shoreline Cleanup. 2018a. Our History. Vancouver (BC): Ocean Wise. [accessed 2019 May 15].

[GCSC] Great Canadian Shoreline Cleanup. 2018b. 2018 Annual Report. Vancouver (BC): Ocean Wise. 15 p.

[GCSC] Great Canadian Shoreline Cleanup. [date unknown]. Site Coordinator Guide. Vancouver (BC): Ocean Wise. 17 p.

Gerdes Z, Ogonowski M, Nybom I, Ek C, Adolfsson-Erici M, Barth A, Gorokhova E. 2019. Microplastic-mediated transport of PCBs? A depuration study with Daphnia magna. PLoS One. 14(2):e0205378.

[GESAMP] Joint Group of Experts on the Scientific Aspects. 2016. Sources, fate and effects of microplastics in the marine environment: Part two of a global assessment. In: Kershaw PJ, Rochman CM, editors. IMO/FAO/UNESCO-IOC/UNIDO/WMO/IAEA/UN/UNEP/UNDP GESAMP, Reports and Studies No. 93. London, England: International Maritime Organization. 220 p. 

Gewert B, Plassmann MM, MacLeod M. 2015. Pathways for degradation of plastic polymers floating in the marine environment. Environ Sci Process Impacts. 17(9):1513-1521.

Geyer R, Jambeck JR, Law KL. 2017. Production, use, and fate of all plastics ever made. Sci Adv. 3(7):e1700782.

Gies EA, LeNoble JL, Noël M, Etemadifar A, Bishay F, Hall ER, Ross PS. 2018. Retention of microplastics in a major secondary wastewater treatment plant in Vancouver, Canada. Mar Pollut Bull. 133:553-561.

Godoy V, Blázquez G, Calero M, Quesada L, Martín-Lara MA. 2019. The potential of microplastics as carriers of metals. Environ Pollut. 255(Pt 3):113363.

Gong M, Yang G, Zhuang L, Zeng EY. 2019. Microbial biofilm formation and community structure on low-density polyethylene microparticles in lake water microcosms. Environ Pollut. 252(Pt A):94-102.

Good TP, June JA, Etnier M, Broadhurst G. 2007. Quantifying the impact of derelict fishing gear on the marine fauna of Puget Sound and the Northwest Straits. In: Proceedings of the International Council for the Exploration of the Sea (ICES) Annual Science Conference; 2007 Sep 17-21; Helsinki (FI): ICES.

Goodman AJ, Walker TR, Brown CJ, Wilson BR, Gazzola V, Sameoto JA. 2020. Benthic marine debris in the Bay of Fundy, eastern Canada: Spatial distribution and categorization using seafloor video footage. Mar Pollut Bull. 150:110722.

Gouin T. 2020. Toward an improved understanding of the ingestion and trophic transfer of microplastic particles: Critical review and implications for future research. Environ Toxicol Chem. 39(6):1119-1137.

Gouin T, Becker RA, Collot AG, Davis JW, Howard B, Inawaka K, Lampi M, Ramon BS, Shi J, Hopp PW. 2019. Towards the development and application of an environmental risk assessment framework for microplastic. Environ Toxicol Chem. 38(10):2087-2100.

Grbić J, Helm P, Athey S, Rochman CM. 2020. Microplastics entering northwestern Lake Ontario are diverse and linked to urban sources. Water Res. 174:115623.

Grigorakis S, Drouillard KG. 2018. Effect of microplastic amendment to food on diet assimilation efficiencies of PCBs by fish. Environ Sci Technol. 52(18):10796-10802. 

Grigorakis S, Mason SA, Drouillard KG. 2017. Determination of the gut retention of plastic microbeads and microfibers in goldfish (Carassius auratus). Chemosphere. 169:233-238.

Gündoğdu S. 2018. Contamination of table salts from Turkey with microplastics. Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 35(5):1006-1014. 

Guo X, Wang J. 2019a. The chemical behaviours of microplastics in marine environment: A review. Mar Pollut Bull. 142:1-14. 

Guo X, Wang J. 2019b. Sorption of antibiotics onto aged microplastics in freshwater and seawater. Mar Pollut Bull. 149:110511.

Hahladakis JN, Velis CA, Weber R, Iacovidou E, Purnell P. 2018. An overview of chemical additives present in plastics: Migration, release, fate and environmental impact during their use, disposal and recycling. J Hazard Mater. 344:179-199. [cited in WHO 2019].

Hall-Stoodley L, Costerton JW, Stoodley P. 2004. Bacterial biofilms: From the natural environment to infectious diseases. Nat Rev Microbiol. 2(2):95-108.

Hallanger IG, Gabrielsen GW. 2018. Plastic in the European Arctic. Brief Report. Tromsø (NO): Norwegian Polar Institute. 23 p.

Hämer J, Gutow L, Köhler A, Saborowski R. 2014. Fate of microplastics in the marine isopod Idotea emarginata. Environ Sci Technol. 48:13451-13458.

Hammer S, Nager RG, Johnson PCD, Furness RW, Provencher JF. 2016. Plastic debris in great skua (Stercorarius skua) pellets corresponds to seabird prey species. Mar Pollut Bull. 103(1-2):206-210.

Hanachi P, Karbalaei S, Walker TR, Cole M, Hosseini SV. 2019. Abundance and properties of microplastics found in commercial fish meal and cultured common carp (Cyprinus carpio). Environ Sci Pollut Res Int. 26(23):23777-23787.

Hankins C, Duffy A, Drisco K. 2018. Scleractinian coral microplastic ingestion: Potential calcification effects, size limits, and retention. Mar Pollut Bull. 135:587-593.

Hansen E, Nilsson NH, Lithner D, Lassen C. 2013. Hazardous substances in plastic materials. Vejle (DK): Norwegian Climate and Pollution Agency. 151 p. Report No. TA 3017. [cited in WHO 2019].

Hantoro I, Löhr AJ, Van Belleghem FGAJ, Widianarko B, Ragas AMJ. 2019. Microplastics in coastal areas and seafood: implications for food safety. Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 36(5):674-711. 

Hara J, Frias J, Nash R. 2020. Quantification of microplastic ingestion by the decapod crustacean Nephrops norvegicus from Irish waters. Mar Pollut Bull. 152:110905.

Harrison JP, Schratzberger M, Sapp M, Osborn AM. 2014. Rapid bacterial colonization of low-density polyethylene microplastics in coastal sediment microcosms. BMC Microbiol. 14:232.

Harrison JP, Hoellein TJ, Sapp M, Tagg AS, Ju-Nam Y, Ojeda JJ. 2018. Microplastic-associated biofilms: a comparison of freshwater and marine environments. Handb Environ Chem. 58:181-201.

Helm PA. 2017. Improving microplastics source apportionment: a role for microplastic morphology and taxonomy? Anal Methods. 9:1328-1331.

Helmroth E, Rijk R, Dekker M, Jongen W. 2002. Predictive modelling of migration from packaging materials into food products for regulatory purposes. Trends Food Sci Technol. 13(3):102-109. [cited in Galloway 2015].

Hendrickson E, Minor EC, Schreiner K. 2018. Microplastic abundance and composition in Western Lake Superior as determined via microscopy, Pyr-GC/MS, and FTIR. Environ Sci Technol. 52(4):1787-1796.

Henry B, Laitala K, Klepp IG. 2019. Microfibres from apparel and home textiles: Prospects for including microplastics in environmental sustainability assessment. Sci Total Environ. 652:483-494. 

Hermabessiere L, Paul-Pont I, Cassone AL, Himber C, Receveur J, Jezequel R, Rakwe ME, Rinnert E, Rivière G, Lambert C, et al. 2019. Microplastic contamination and pollutant levels in mussels and cockles collected along the channel coasts. Environ Pollut. 250:807-819.

Hermsen E, Mintenig SM, Besseling E, Koelmans AA. 2018. Quality criteria for the analysis of microplastic in biota samples: A critical review. Environ Sci Technol. 52(18):10230-10240. 

Hernandez LM, Xu EG, Larsson HCE, Tahara R, Maisuria VB, Tufenkji N. 2019. Plastic Teabags Release Billions of Microparticles and Nanoparticles into Tea. Environ Sci Technol. 53(21):12300-12310.

Hernandez-Milian G, Lusher A, MacGabban S, Rogan E. 2019. Microplastics in grey seal (Halichoerus grypus) intestines: Are they associated with parasite aggregations?. Mar Pollut Bull. 146:349-354.

Hesterberg TW, McConnell EE, Miller WC, Hamilton R, Bunn WB. 1992. Pulmonary toxicity of inhaled polypropylene fibers in rats. Fundam Appl Toxicol. 19(3):358-366.

Hidalgo-Ruz V, Gutow L, Thompson RC, Thiel M. 2012. Microplastics in the marine environment: a review of the methods used for identification and quantification. Environ Sci Tech. 46(6):3060-3075. 

Hipfner JM, Galbraith M, Tucker S, Studholme KR, Domalik AD, Pearson SF, Good TP, Ross PS, Hodum P. 2018. Two forage fishes as potential conduits for the vertical transfer of microfibers in Northeastern Pacific Ocean food webs. Environ Pollut. 239:215-222.

Hoellein T, Rojas M, Pink A, Gasior J, Kelly J. 2014. Anthropogenic litter in urban freshwater ecosystems: Distribution and microbial interactions. PLoS One. 9(6):e98485.

Horn D, Miller M, Anderson S, Steele C. 2019. Microplastics are ubiquitous on California beaches and enter the coastal food web through consumption by Pacific mole crabs. Mar Pollut Bull. 139:231-237.

Horton AA, Dixon SJ. 2017. Microplastics: An introduction to environmental transport processes. WIREs Water. 5(2):e1268. 

Horton AA, Newbold LK, Palacio-Cortés AM, Spurgeon DJ, Pereira MG, Carter H, Gweon HS, Vijver MG, van Bodegom PM, Navarro da Silva MA, et al. 2020. Accumulation of polybrominated ethers and microbiome response in the great pond snail Lymnaea stagnalis with exposure to nylon (polyamide) microplastics. Ecotoxicol Environ Saf. 188:109882.

Hossain MS, Rahman MS, Uddin MN, Sharifuzzaman SM, Chowdhury SR, Sarker S, Nawaz Chowdhury MS. 2020. Microplastic contamination in Penaeid shrimp from the Northern Bay of Bengal. Chemosphere. 238:124688.

Huerta Lwanga E, Gertsen H, Gooren H, Peters P, Salánki T, van der Ploeg M, Besseling E, Koelmans AA, Geissen V. 2016. Microplastics in the terrestrial ecosystem: Implications for Lumbricus terrestris (Oligochaeta, Lumbricidae). Environ Sci Technol. 50(5):2685-2691. 

Huerta Lwanga E, Gertsen H, Gooren H, Peters P, Salánki T, van der Ploeg M, Besseling E, Koelmans AA, Geissen V. 2017. Incorporation of microplastics from litter into burrows of Lumbricus terrestris. Environ Pollut. 220(Pt A):523-531. 

Huerta Lwanga E, Thapa B, Yang X, Gertsen H, Salánka T, Geissen V, Garbeva P. 2018. Decay of low-density polyethylene by bacteria extracted from earthworm's guts: A potential for soil restoration. Sci Total Environ. 624:753-757.

Hurley RR, Nizzetto L. 2018. Fate and occurrence of micro(nano)plastics in soils: Knowledge gaps and possible risks. Curr Opin Environ Sci Health. 1:6-11. 

Hurley RR, Woodward J, Rothwell JJ. 2018. Microplastic contamination of river beds significantly reduced by catchment-wide flooding. Nat Geosci. 11:251-257.

Hussain N, Jaitley V, Florence AT. 2001. Recent advances in the understanding of uptake of microparticulates across the gastrointestinal lymphatics. Adv Drug Deliv Rev. 50(1-2):107-142. 

Imran M, Das KR, Naik MM. 2019. Co-selection of multi-antibiotic resistance in bacterial pathogens in metal and microplastic contaminated environments: An emerging health threat. Chemosphere. 215:846-857.

Iñiguez ME, Conesa JA, Fullana A. 2017. Microplastics in Spanish Table Salt. Sci Rep. 7(1):8620.

Jabeen K, Li B, Chen Q, Su L, Wu C, Hollert H, Shi H. 2018. Effects of virgin microplastics on goldfish (Carassius auratus). Chemosphere. 213:323-332.

Jacob H, Gilson A, Lanctôt C, Besson M, Metian M, Lecchini D. 2019. No effect of polystyrene microplastics on foraging activity and survival in a post-larvae coral-reef fish, Acanthurus triostegus. Bull Environ Contam Toxicol. 102(4):457-461. 

Jacobsen JK, Massey L, Gulland F. 2010. Fatal ingestion of floating net debris by two sperm whales (Physeter macrocephalus). Mar Pollut Bull. 60(5):765-767. 

Jahnke A, Arp HPH, Escher BI, Gewert B, Gorokhova E, Kühnel D, Ogonowski M, Potthoff A, Rummel C, Schmitt-Jansen M, et al. 2017. Reducing uncertainty and confronting ignorance about the possible impacts of weathering plastic in the marine environment. Environ Sci Technol Lett. 4(3):85-90. [cited in WHO 2019].

Jaikumar G, Baas J, Brun NR, Vijver MG, Bosker T. 2018. Acute sensitivity of three Cladoceran species to different types of microplastics in combination with thermal stress. Environ Pollut. 239:733-740.

Jarvis P, Jefferson B, Gregory J, Parsons SA. 2005. A review of floc strength and breakage. Water Res. 39(14):3121-3137.

Jeong CB, Won EJ, Kang HM, Lee MC, Hwang DS, Hwang UK, Zhou B, Souissi S, Lee SJ, Lee JS. 2016. Microplastic size-dependent toxicity, oxidative stress induction, and p-JNK and p-p38 activation in the monogonont rotifer (Brachionus koreanus). Environ Sci Technol. 50(16):8849-8857. 

Jeong CB, Kang HM, Lee MC, Kim DH, Han J, Hwang DS, Souissi S, Lee SJ, Shin KH, Park HG, et al. 2017. Adverse effects of microplastics and oxidative stress-induced MAPK/Nrf2 pathway-mediated defense mechanisms in the marine copepod Paracyclopina nana. Sci Rep. 7:41323. 

Jepsen EM, de Bruyn PJN. 2019. Pinniped entanglement in oceanic plastic pollution: A global review. Mar Pollut Bull. 145:295-305.

Jin Y, Lu L, Tu W, Luo T, Fu Z. 2019. Impacts of polystyrene microplastic on the gut barrier, microbiota and metabolism of mice. Sci Total Environ. 649:308-317.

Johansen MP, Cresswell T, Davis J, Howard DL, Howell NR, Prentice E. 2019. Biofilm-enhanced adsorption of strong and weak cations onto different microplastic sample types: Use of spectroscopy, microscopy and radiotracer methods. Wat Res. 158:392-400.

Jones KL, Hartl MGJ, Bell MC, Capper A. 2020. Microplastic accumulation in a Zostera marina L. bed at Deerness Sound, Orkney, Scotland. Mar Pollut Bull. 152:110883.

Jovanović B. 2017. Ingestion of microplastics by fish and its potential consequences from a physical perspective. Integr Environ Assess Manag. 13(3):510-515.

Jovanović B, Gökdağ K, Güven O, Emre Y, Whitley EM, Kideys AE. 2018. Virgin microplastics are not causing imminent harm to fish after dietary exposure. Mar Pollut Bull. 130:123-131.

Ju H, Zhu D, Qiao M. 2019. Effects of polyethylene microplastics on the gut microbial community, reproduction and avoidance behaviors of the soil springtail, Folsomia candida. Environ Pollut. 247:890-897.

Julienne F, Delorme N, Lagarde F. 2019. From macroplastics to microplastics: role of water in the fragmentation of polyethylene. Chemosphere. 236:124409.

Kaiser D, Kowalski N, Waniek JJ. 2017. Effects of biofouling on the sinking behaviour of microplastics. Environ Res Lett. 12(12):124003. 

Kanhai LDK, Gårdfeldtc K, Lyashevska O, Hassellöv M, Thompson RC, O'Connor I. 2018. Microplastics in sub-surface waters of the Arctic Central Basin. Mar Pollut Bull. 130:8-18.

Karami A, Golieskardi A, Ho YB, Larat V, Salamatinia B. 2017a. Microplastics in eviscerated flesh and excised organs of dried fish. Sci Rep. 7(1):5473.

Karami A, Golieskardi A, Keong Choo C, Larat V, Galloway TS, Salamatinia B. 2017b. The presence of microplastics in commercial salts from different countries. Sci Rep. 7:46173.

Karami A, Golieskardi A, Choo CK, Larat V, Karbalaei S, Salamatinia B. 2018. Microplastic and mesoplastic contamination in canned sardines and sprats. Sci Total Environ. 612:1380-1386.

Karbalaei S, Golieskardi A, Watt DU, Boiret M, Hanachi P, Walker TR, Karami A. 2020. Analysis and inorganic composition of microplastics in commercial Malaysian fish meals. Mar Pollut Bull. 150:110687.

Karlsson TM, Vethaak AD, Almroth BC, Ariese F, van Velzen M, Hassellöv M, Leslie HA. 2017. Screening for microplastics in sediment, water, marine invertebrates and fish: method development and microplastic accumulation. Mar Pollut Bull. 122(1-2):403-408.

Karthik R, Robin RS, Purvaja R, Ganguly D, Anandavelu I, Raghuraman R, Hariharan G, Ramakrishna A, Ramesh R. 2018. Microplastics along the beaches of southeast coast of India. Sci Total Environ. 645:1388-1399.

Katsanevakis S, Verriopoulos G, Nicolaidou A, Thessalou-Legaki M. 2007. Effect of marine litter on the benthic megafauna of coastal soft bottoms: A manipulative field experiment. Mar Pollut Bull. 54(6):771-778. 

Kay P, Hiscoe R, Moberley I, Bajic L, McKenna N. 2018. Wastewater treatment plants as a source of microplastics in river catchments. Environ Sci Pollut Res Int. 25(20):20264-20267.

Kazmiruk TN, Kazmiruk VD, Bendell LI. 2018. Abundance and distribution of microplastics within surface sediments of a key shellfish growing region of Canada. PLoS One. 13(5):e0196005.

Kedzierski M, Lechat B, Sire O, Le Maguer G, Le Tilly V, Bruzaud S. 2020. Microplastic contamination of packaged meat: Occurrence and associated risks. Food Packag Shelf Life. 24:100489.

Keswani A, Oliver DM, Gutierrez T, Quilliam RS. 2016. Microbial hitchhikers on marine plastic debris: Human exposure risks at bathing waters and beach environments. Mar Environ Res. 118:10-19.

Kettner MT, Rojas-Jimenez K, Oberbeckmann S, Labrenz M, Grossart H. 2017. Microplastics alter composition of fungal communities in aquatic ecosystems. Environ Microbiol. 19(11):4447-4459.

Kettner MT. 2018. Microbial colonization of microplastic particles in aquatic systems [dissertation]. Potsdam (DE): University of Potsdam.

Kienitz AT. 2013. Marine Debris in the Coastal Environment of Iceland’s Nature Reserve, Hornstrandir - Sources, Consequences and Prevention Measures [Master's Thesis]. Ísafjörður (IS): University of Akureyri. [cited in PAME 2019].

Kim JS, Lee HJ, Kim SK, Kim HJ. 2018. Global pattern of microplastics (MPs) in commercial food-grade salts: Sea salt as an indicator of seawater MP pollution. Environ Sci Technol. 52(21):12819-12828.

Kim SW, An YJ. 2019. Soil microplastics inhibit the movement of springtail species. Environ Int. 126:699-706.

Kirstein IV, Kirmizi S, Wichels A, Garin-Fernandez A, Erler R, Löder M, Gerdts G. 2016. Dangerous hitchhikers? Evidence for potentially pathogenic Vibrio spp. on microplastic particles. Mar Environ Res. 120:1-8. 

Klein M, Fischer EK. 2019. Microplastic abundance in atmospheric deposition within the Metropolitan area of Hamburg, Germany. Sci Total Environ. 685:96-103. 

Koelmans AA, Besseling E, Shim WJ. 2015. Nanoplastics in the aquatic environment: Critical review. In: Bergmann M, Gutow L, Klages M, editors. Marine Anthropogenic Litter. Cham (CH): Springer. p. 325-340. 

Koelmans AA, Bakir A, Burton GA, Janssen CR. 2016. Microplastics as a vector for chemicals in the aquatic environment: Critical review and model-supported reinterpretation of empirical studies. Environ Sci Technol. 50:3315-3326. 

Koelmans AA, Mohamed Nor NH, Hermsen E, Kooi M, Mintenig SM, De France J. 2019. Microplastics in freshwaters and drinking water: Critical review and assessment of data quality. Water Res. 155:410-422. 

Kole PJ, Löhr AJ, Van Belleghem FGAJ, Ragas AMJ. 2017. Wear and tear of tyres: a stealthy source of microplastics in the environment. Int J Environ Res Public Health. 14(10):1265.

Kooi M, Koelmans AA. 2019. Simplifying microplastic via continuous probability distributions for size, shape, and density. Environ Sci Technol Lett. 6(9):551−557.

Korez S. Gutow L, Saborowski R. 2019. Feeding and digestion of the marine isopod Idotea emarginata challenged by poor food quality and microplastics. Comp Biochem Physiol C Toxicol Pharmacol. 226:108586.

Kosuth M, Mason SA, Wattenberg EV. 2018. Anthropogenic contamination of tap water, beer, and sea salt. PLoS One. 13(4):e0194970.

Kratina P, Watts TJ, Green DS, Kordas RL, O'Gorman EJ. 2019. Interactive effects of warming and microplastics on metabolism but not feeding rates of a key freshwater detritivore. Environ Pollut. 255(Pt 2):113259. 

Kreider ML, Unice KM, Panko JM. 2019. Human health risk assessment of tire and road wear particles (TRWP) in air. Hum Ecol Risk Assess. In press.

Kühn S, Bravo Rebolledo EL, van Franeker JA. 2015. Deleterious effects of litter on marine life. In: Bergmann M, Gutow L, Klages M, editors. Marine Anthropogenic Litter. Cham (CH): Springer. Chapter 4. p. 75-116. 

Kumar VE, Ravikumar G, Jeyasanta KI. 2018. Occurrence of microplastics in fishes from two landing sites in Tuticorin, south east coast of India. Mar Pollut Bull. 135:889-894. 

Kutralam-Muniasamy G, Pérez-Guevara F, Elizalde-Martínez I, Shruti VC. 2020. Branded milks - Are they immune from microplastics contamination? Sci Total Environ. 714:136823.

Lachenmeier DW, Kocareva J, Noack D, Kuballa T. 2015. Microplastic identification in German beer - an artefact of laboratory contamination? Dtsch Lebensmitt Rundsch. 111(10):437-440.

Laganà P, Caruso G, Corsi I, Bergami E, Venuti V, Majolino D, La Ferla R, Azzaro M, Cappello S. 2019. Do plastics serve as a possible vector for the spread of antibiotic resistance? First insights from bacteria associated to a polystyrene piece from King George Island (Antarctica). Int J Hyg Environ Health. 222(1):89-100.

Lahens L, Strady E, Kieu-Le TC, Dris R, Boukerma K, Rinnert E, Gasperi J, Tassin B. 2018. Macroplastic and microplastic contamination assessment of a tropical river (Saigon River, Vietnam) transversed by a developing megacity. Environ Pollut. 236:661-671.

Lahive E, Walton A, Horton AA, Spurgeon DJ, Svendsen C. 2019. Microplastic particles reduce reproduction in the terrestrial worm Enchytraeus crypticus in a soil exposure. Environ Pollut. 255(Pt 2):113174.

Lamb JB, Willis BL, Fiorenza EA, Couch CS, Howard R, Rader DN, True JD, Kelly LA, Ahmad A, Jompa J, et al. 2018. Plastic waste associated with disease on coral reefs. Science. 359(6374):460-462.

Lambert S, Wagner M. 2017. Environmental performance of bio-based biodegradable plastics: the road ahead. Chem Soc Rev. 46(22):6855-6871.

Lares M, Ncibi MC, Sillanpaa Ma, Sillanpaa Mi. 2018. Occurrence, identification and removal of microplastic particles and fibers in conventional activated sludge process and advanced MBR technology. Water Res. 133:236-246.

Laskin S, Drew RT, Cappiello VP, Kuschner M. 1972. Inhalation studies with freshly generated polyurethane foam dust. In: Mercer TT, Morrow PE, Stober W, editors. Assessment of Airborne Particles. Springfield (IL): Charles C. Thomas. p. 382-404.

Lavers JL, Bond AL. 2016. Ingested plastic as a route for trace metals in Laysan Albatross (Phoebastria immutabilis) and Bonin Petrel (Pterodroma hypoleuca) from Midway Atoll. Mar Pollut Bull. 110:493-500. 

Lavers JL, Hutton I, Bond AL. 2019. Clinical pathology of plastic ingestion in marine birds and relationships with blood chemistry. Environ Sci Technol. 53(15):9224-9231.

Lebreton L, Slat B, Ferrari F, Sainte-Rose B, Aitken J, Marthouse R, Hajbane S, Cunsolo S, Schwarz A, Levivier A, et al. 2018. Evidence that the Great Pacific Garbage Patch is rapidly accumulating plastic. Sci Rep. 8(1):4666.

Lee H, Kunz A, Shim WJ, Walther BA. 2019. Microplastic contamination of table salts from Taiwan, including a global review. Sci Rep. 9(1):10145.

Lee KW, Shim WJ, Kwon OY, Kang JH. 2013. Size-dependent effects of micro polystyrene particles in the marine copepod Tigriopus japonicus. Environ Sci Technol. 47(19):11278-11283.

Lei L, Liu M, Song Y, Lu S, Hu J, Cao C, Xie B, Shi H, He D. 2018b. Polystyrene (nano)microplastics cause size-dependent neurotoxicity, oxidative damage and other adverse effects in Caenorhabditis elegans. Environ Sci Nano. 5:2009-2020.

Lei L, Wu S, Lu S, Liu M, Song Y, Fu Z, Shi H, Raley-Susman KM, He D. 2018a. Microplastic particles cause intestinal damage and other adverse effects in zebrafish Danio rerio and nematode Caenorhabditis elegans. Sci Total Environ. 619-620:1-8.

Leonas KK, Gorden RW. 1993. An accelerated laboratory study evaluating the disintegration rates of plastic films in simulated aquatic environments. J Environ Polym Degrad. 1:45-51.

Letcher RJ, Bustnes JO, Dietz R, Jenssen BM, Jørgensen EH, Sonne C, Verreault J, Vijayan MM, Gabrielsen GW. 2010. Exposure and effects assessment of persistent organohalogen contaminants in arctic wildlife and fish. Sci Total Environ. 408(15):2995-3043.

Leung J, Chan KYK. 2018. Microplastics reduced posterior segment regeneration rate of the polychaete Perinereis aibuhitensis. Mar Pollut Bull. 129(2):782-786.

Li B, Ding Y, Cheng X, Sheng D, Xu Z, Rong Q, Wu Y, Zhao H, Ji X, Zhang Y. 2020. Polyethylene microplastics affect the distribution of gut microbiota and inflammation development in mice. Chemosphere. 244:125492.

Li J, Yang D, Li L, Jabeen K, Shi H. 2015. Microplastics in commercial bivalves from China. Environ Pollut. 207:190-195.

Li J, Green C, Reynolds A, Shi H, Rotchell JM. 2018a. Microplastics in mussels sampled from coastal waters and supermarkets in the United Kingdom. Environ Pollut. 241:35-44.

Li J, Liu H, Chen JP. 2018b. Microplastics in freshwater systems: A review on occurrence, environmental effects, and methods for microplastics detection. Water Res. 137:362-374.

Li N, Georas S, Alexis N, Fritz P, Xia T, Williams MA, Horner E, Nel A. 2016. A work group report on ultrafine particles (American Academy of Allergy, Asthma & Immunology): why ambient ultrafine and engineered nanoparticles should receive special attention for possible health outcomes in human subjects. J Allergy Clin Immunol. 138(2):386-396. 

Li Q, Feng Z, Zhang T, Ma C, Shi H. 2020. Microplastics in the commercial seaweed nori. J Hazard Mater. 388:122060. 

Li WC, Tse HF, Fok L. 2016. Plastic waste in the marine environment: A review of sources, occurrence and effects. Sci Total Environ. 566-567:333-349.

Li Y, Li M, Li Z, Yang L, Liu X. 2019. Effects of particle size and solution chemistry on triclosan sorption on polystyrene microplastic. Chemosphere. 231:308-314.

Li Y, Shao L, Wang W, Zhang M, Feng X, Li W, Zhang D. 2020. Airborne fiber particles: Types, size and concentration observed in Beijing. Sci Total Environ. 705:135967.

Liao Y-L, Yang J-Y. 2020. Microplastic serves as a potential vector for Cr in an in-vitro human digestive model. Sci Total Environ. 703:134805.

Liboiron F, Ammendolia J, Saturno J, Melvin J, Zahara A, Richard N, Liboiron M. 2018. A zero percent plastic ingestion rate by silver hake (Merluccius bilinearis) from the south coast of Newfoundland, Canada. Mar Pollut Bull. 131:267-275. 

Liboiron M, Melvin J, Richard N, Saturno J, Ammendolia J, Liboiron F, Charron L, Mather C. 2019. Low incidence of plastic ingestion among three fish species significant for human consumption on the island of Newfoundland, Canada. Mar Pollut Bull. 141:244-248. 

Liebezeit G, Liebezeit E. 2013. Non-pollen particulates in honey and sugar. Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 30(12):2136-2140.

Liebezeit G, Liebezeit E. 2014. Synthetic particles as contaminants in German beers. Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 31(9):1574-1578.

Liebezeit G, Liebezeit E. 2015. Origin of Synthetic Particles in Honeys. Pol J Food Nutr Sci. 65(2):143-147.

Lithner D, Larsson A, Dave G. 2011. Environmental and health hazard ranking and assessment of plastic polymers based on chemical composition. Sci Total Environ. 409:3309-3324.

Liu C, Li J, Zhang Y, Wang L, Deng J, Gao Y, Yu L, Zhang J, Sun H. 2019. Widespread distribution of PET and PC microplastics in dust in urban China and their estimated human exposure. Environ Int. 128:116-124.

Liu K, Wang X, Fang T, Xu P, Zhu L, Li D. 2019a. Source and potential risk assessment of suspended atmospheric microplastics in Shanghai. Sci Total Environ. 675:462-471.

Liu K, Wang X, Wei N, Song Z, Li D. 2019b. Accurate quantification and transport estimation of suspended atmospheric microplastics in megacities: Implications for human health. Environ Int. 132:105127.

Liu K, Wu T, Wang X, Song Z, Zong C, Wei N, Li D. 2019c. Consistent transport of terrestrial microplastics to the ocean through atmosphere. Environ Sci Technol. 53(18):10612-10619

Liu M, Lu S, Song Y, Lei L, Hu J, Lv W, Zhou W, Cao C, Shi H, Yang X, He D. 2018. Microplastic and mesoplastics pollution in farmland soils in suburbs of Shanghai, China. Environ Pollut. 242:855-862.

Liu S, Gunawan C, Barraud N, Rice SA, Harry EJ, Amal R. 2016. Understanding, monitoring, and controlling biofilm growth in drinking water distribution systems. Environ Sci Technol. 50(17):8954-8976.

Liu X, Xu J, Zhao Y, Shi H, Huang C-H. 2019. Hydrophobic sorption behaviors of 17β-Estradiol on environmental microplastics. Chemosphere. 226:726-735.

Liu Z, Yu P, Cai M, Wu D, Zhang M, Chen M, Zhao Y. 2019. Effects of microplastics on the innate immunity and intestinal microflora of juvenile Eriocheir sinensis. Sci Total Environ. 685:836-846. 

Lo HKA, Chan KYK. 2018. Negative effects of microplastic exposure on growth and development of Crepidula onyx. Environ Pollut. 233:588-595.

Long M, Moriceau B, Gallinari M, Lambert C, Huvet A, Raffray J, Soudant P. 2015. Interactions between microplastics and phytoplankton aggregates: Impact on their respective fates. Mar Chem. 175:39-46.

Lu L, Wan Z, Luo T, Fu Z, Jin Y. 2018. Polystyrene microplastics induce gut microbiota dysbiosis and hepatic lipid metabolism disorder in mice. Sci Total Environ. 631-632:449-458.

Lu Y, Zhang Y, Deng Y, Jiang W, Zhao Y, Geng J, Ding L, Ren H. 2016. Uptake and accumulation of polystyrene microplastics in zebrafish (Danio rerio) and toxic effects in liver. Environ Sci Technol. 50(7):4054-4060.

Luo T, Wang C, Pan Z, Jin C, Fu Z, Jin Y. 2019a. Maternal polystyrene microplastic exposure during gestation and lactation altered metabolic homeostasis in the dams and their F1 and F2 offspring. Environ Sci Technol. 53(18):10978-10992. 

Luo T, Zhang Y, Wang C, Wang X, Zhou J, Shen M, Zhao Y, Fu Z, Jin Y. 2019b. Maternal exposure to different sizes of polystyrene microplastics during gestation causes metabolic disorders in their offspring. Environ Pollut. 255(Pt 1):113122. 

Luo H, Zhao Y, Li Y, Xiang Y, He D, Pan X. 2020. Aging of microplastics affects their surface properties, thermal decomposition, additives leaching and interactions in simulated fluids. Sci Total Environ. 714:136862.

Lusher AL, McHugh M, Thompson RC. 2013. Occurrence of microplastics in the gastrointestinal tract of pelagic and demersal fish from the English Channel. Mar Pollut Bull. 67(1-2):94-99.

Lusher AL, Hernandez-Milian G, O’Brien J, Berrow S, O’Connor I, Officer R. 2015b. Microplastic and macroplastic ingestion by a deep diving, oceanic cetacean: The True’s beaked whale Mesoplodon mirus. Environ Pollut. 199:185-191. 

Lusher AL, Tirelli V, O’Connor I, Officer R. 2015a. Microplastics in Arctic polar waters: the first reported values of particles in surface and sub-surface samples. Sci Rep. 5:14947.

Lusher AL, Welden NA, Sobral P, Cole M. 2017. Sampling, isolating and identifying microplastics ingested by fish and invertebrates. Anal Methods. 9:1346-1360.

Ma B, Xue W, Ding Y, Hu C, Liu H, Qu J. 2018. Removal characteristics of microplastics by Fe-based coagulants during drinking water treatment. J Environ Sci. 78:267-275. 

Ma J, Zhao J, Zhu Z, Li L, Yu F. 2019. Effect of microplastic size on the adsorption behavior and mechanism of triclosan on polyvinyl chloride. Environ Pollut. 254(Pt B):113104.

Ma-Hock L, Landsiedel R, Wiench K, Geiger D, Strauss V, Gröters S, van Ravenzwaay B, Gerst M, Wohlleben W, Scherer G. 2012. Short-term rat inhalation study with aerosols of acrylic ester-based polymer dispersions containing a fraction of nanoparticles. Int J Toxicol. 31(1):46-57. 

Maaß S, Daphi D, Lehmann A, Rillig MC. 2017. Transport of microplastics by two collembolan species. Environ Pollut. 225:456-459. 

Macali A, Semenov A, Venuti V, Crupi V, D’Amico F, Rossi B, Corsi I, Bergami E. 2018. Episodic records of jellyfish ingestion of plastic items reveal a novel pathway for trophic transference of marine litter. Sci Rep. 8(1):6105. 

Maes T, Jessop R, Wellner N, Haupt K, Mayes AG. 2017. A rapid-screening approach to detect and quantify microplastics based on fluorescent tagging with Nile Red. Sci Rep. 7:44501. 

Magara G, Elia AC, Syberg K, Khan FR. 2018. Single contaminant and combined exposures of polyethylene microplastics and fluoranthene: accumulation and oxidative stress response in the blue mussel, Mytilus edulis. J Toxicol Environ Health A. 81(16):761-773. 

Magnusson K, Norén F. 2014. Screening of microplastic particles in and down-stream a wastewater treatment plant [PDF]. Stockholm (SE): IVL Swedish Environmental Research Institute. [accessed 2019 May 29].

Magnusson K, Eliasson K, Fråne A, Haikonen K, Hultén J, Olshammar M, Stadmark J, Voisin A. 2016. Swedish sources and pathways for microplastics to the marine environment. A review of existing data. IVL - Swedish Environmental Research Institute.

Mahler GJ, Esch MB, Tako E, Southard TL, Archer SD, Glahn RP, Shuler ML. 2012. Oral exposure to polystyrene nanoparticles affects iron absorption. Nat Nanotechnol. 7(4):264-271.

Mahon AM, O’Connell B, Healy MG, O’Connor I, Officer R, Nas R, Morrison L. 2017. Microplastics in Sewage Sludge: Effects of Treatment. Environ Sci Technol. 51(2):810-818. 

Malafaia G, de Souza AM, Pereira AC, Gonçalves S, da Costa Araújo AP, Ribeiro RX, Rocha TL. 2020. Developmental toxicity in zebrafish exposed to polyethylene microplastics under static and semi-static aquatic systems. Sci Total Environ. 700:134867.

Mani T, Hauk A, Walter U, Burkhardt-Holm P. 2015. Microplastics profile along the Rhine River. Sci Rep. 5:17988).

Mao Y, Ai H, Chen Y, Zhang Z, Zeng P, Kang L, Li W, Gu W, He Q, Li H. 2018. Phytoplankton response to polystyrene microplastics: Perspective from an entire growth period. Chemosphere. 208:59-68.

Martins A, Guilhermino L. 2018. Transgenerational effects and recovery of microplastics exposure in model populations of the freshwater cladoceran Daphnia magna Straus. Sci Total Environ. 631-632:421-428. 

Mason SA, Kammin L, Eriksen M, Aleid G, Wilson S, Box C, Williamson N, Riley A. 2016. Pelagic plastic pollution within surface waters of Lake Michigan, USA. J Great Lakes Res. 42:753-759.

Mason SA, Welch VG, Neratko J. 2018. Synthetic Polymer Contamination in Bottled Water. Front Chem. 6:407.

Mason SA, Daily J, Aleid G, Ricotta R, Smith M, Donnelly K, Knauff R, Edwards W, Hoffman MJ. 2020. High levels of pelagic plastic pollution within the surface waters of Lakes Erie and Ontario. J Great Lakes Res. 46(2):277-288.

Mateos-Cárdenas A, Scott DT, Seitmaganbetova G, van Pelt FNAM, O’Halloran J, Jansen MAK. 2019. Polyethylene microplastics adhere to Lemna minor (L.), yet have no effects on plant growth or feeding by Gammarus duebeni (Lillj.). Sci Total Environ. 689:413-421.

Mathalon A, Hill P. 2014. Microplastic fibers in the intertidal ecosystem surrounding Halifax Harbor, Nova Scotia. Mar Pollut Bull. 81(1):69-79.

Mato Y, Isobe T, Takada H, Kanehiro H, Ohtake C, Kaminuma T. 2001. Plastic resin pellets as a transport medium for toxic chemicals in the marine environment. Environ Sci Technol. 35:318-324.

Mazurais D, Ernande B, Quazuguel P, Severe A, Huelvan C, Madec L, Mouchel O, Soudant P, Robbens J, Huvet A, et al. 2015. Evaluation of the impact of polyethylene microbeads ingestion in European sea bass (Dicentrarchus labrax) larvae. Mar Environ Res. 112:78-85.

McCormick AR, Hoellein TJ, Mason SA, Schluep J, Kelly JJ. 2014. Microplastic is an abundant and distinct microbial habitat in an urban river. Environ Sci Technol. 48(20):11863-11871. 

McCormick AR, Hoellein TJ, London MG, Hittie J, Scott JW, Kelly JJ. 2016. Microplastic in surface waters of urban rivers: concentration, sources, and associated bacterial assemblages. Ecosphere. 7(11):e01556. 

McWilliams M, Liboiron M, Wiersma Y. 2018. Rocky shoreline protocols miss microplastics in marine debris surveys (Fogo Island, Newfoundland and Labrador). Mar Pollut Bull. 129:480-486.

Menezes R, da Cunha-Neto MA, de Mesquita GC, da Silva GB. 2019. Ingestion of macroplastic debris by the common dolphinfish (Coryphaena hippurus) in the Western Equatorial Atlantic. Mar Pollut Bull. 141:161-163. 

Menicagli V, Balestri E, Lardicci C. 2019a. Exposure of coastal dune vegetation to plastic bag leachates: A neglected impact of plastic litter. Sci Total Environ. 683:737-748.

Menicagli V, Balestri E, Vallerini F, Castelli A, Lardicci C. 2019b. Adverse effects of non-biodegradable and compostable plastic bags on the establishment of coastal dune vegetation: First experimental evidences. Environ Pollut. 252(Pt 2):188-195.

Merski JA, Johnson WD, Muzzio M, Lyang NL, Gaworski CL. 2008. Oral toxicity and bacterial mutagenicity studies with a spunbond polyethylene and polyethylene terephthalate polymer fabric. Int J Toxicol. 27(5):387-395.

Mikoczy Z, Welinder H, Tinnerberg H, Hagmar L. 2004. Cancer incidence and mortality of isocyanate exposed workers from the Swedish polyurethane foam industry: updated findings 1959–98. Occup Environ Med. 61(5):432-437.

Mintenig SM, Loder MGJ, Primpke S, Gerdts G. 2019. Low numbers of microplastics detected in drinking water from ground water sources. Sci Total Environ. 648:631-635.

Mohamed Nor NH, Koelmans AA. 2019. Transfer of PCBs from microplastic under simulated gut fluid conditions is biphasic and reversible. Environ Sci Technol. 53(4):1874-1883.

Mohsen M, Zhang L, Sun L, Lin C, Wang Q, Yang H. 2020. Microplastic fibers transfer from the water to the internal fluid of the sea cucumber Apostichopus japonicus. Environ Pollut. 257:113606.

Moore RC, Loseto L, Noel M, Etemadifar A, Brewster JD, MacPhee S, Bendell L, Ross PS. 2020a. Microplastics in beluga whales (Delphinapterus leucas) from the Eastern Beaufort Sea. Mar Pollut Bull. 150:110723.

Moore RE, Millar BC, Moore JE. 2020b. Antimicrobial resistance (AMR) and marine plastics: Can food packaging litter act as a dispersal mechanism for AMR in oceanic environments? Mar Pollut Bull. 150:110702.

Mouchi V, Chapron L, Peru E, Pruski AM, Meistertzheim A-L, Vétion G, Galand PE, Lartaud F. 2019. Long-term aquaria study suggests species-specific responses of two cold-water corals to macro-and microplastics exposure. Environ Pollut. 253:322-329.

Mühlschlegel P, Hauk A, Walter U, Sieber R. 2017. Lack of evidence for microplastic contamination in honey. Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 34(11):1982-1989. 

Munier B, Bendell LI. 2018. Macro and micro plastics sorb and desorb metals and act as a point source of trace metals to coastal ecosystems. PLoS One. 13(2):e0191759.

Muncke J. 2011. Endocrine disrupting chemicals and other substances of concern in food contact materials: an updated review of exposure, effect and risk assessment. J Steroid Biochem Mol Biol. 127(1-2):118-127. [cited in Galloway 2015].

Murphy F, Ewins C, Carbonnier F, Quinn B. 2016. Wastewater treatment works (WwTW) as a source of microplastics in the aquatic environment. Environ Sci Technol. 50(11):5800-5808.

Murphy F, Russell M, Ewins C, Quinn B. 2017. The uptake of macroplastic & microplastic by demersal & pelagic fish in the Northeast Atlantic around Scotland. Mar Pollut Bull. 122(1-2):353-359. 

Murray F, Cowie PR. 2011. Plastic contamination in the decapod crustacean Nephrops norvegicus (Linnaeus, 1758). Mar Pollut Bull. 62(6):1207-1217.

Naik RK, Naik MM, D'Costa PM, Shaikh F. 2019. Microplastics in ballast water as an emerging source and vector for harmful chemicals, antibiotics, metals, bacterial pathogens and HAB species: A potential risk to the marine environment and human health. Mar Pollut Bull. 149:110525.

Naji A, Nuri M, Vethaak AD. 2018. Microplastics contamination in molluscs from the northern part of the Persian Gulf. Environ Pollut. 235:113-120.

Napper IE, Bakir A, Rowland SJ, Thompson RC. 2015. Characterisation, quantity and sorptive properties of microplastics extracted from cosmetics. Mar Pollut Bull. 99:178-85.

Napper IE, Thompson RC. 2019. Environmental deterioration of biodegradable, oxo-biodegradable, compostable, and conventional plastic carrier bags in the sea, soil, and open-air over a 3-year period. Environ Sci Technol. 53(9):4775-4783.

Ng EL, Huerta Lwanga E, Eldridge SM, Johnston P, Hu HW, Geissen V, Chen D. 2018. An overview of microplastic and nanoplastic pollution in agroecosystems. Sci Total Environ. 627:1377-1388.

Nguyen B, Claveau-Mallet D, Hernandez LM, Xu EG, Farner JM, Tufenkji N. 2019. Separation and analysis of microplastics and nanoplastics in complex environmental samples. Acc Chem Res. 52:858-866.

Noble WC, Lidwell OM, Kingston D. 1963. The size distribution of airborne particles carrying micro-organisms. J Hyg. 61(4):385-391. [cited in Prata 2018].

Nizzetto L, Futter M, Langaas S. 2016. Are agricultural soils dumps for microplastics of urban origin? Environ Sci Technol. 50:10777-10779. 

Novotna K, Cermakova L, Pivokonska L, Cajthaml T, Pivokonsky M. 2019. Microplastics in drinking water treatment- Current knowledge and research needs. Sci Total Environ. 667:730-740. 

Obbard RW, Sadri S, Wong YQ, Khitun AA, Baker I, Thompson RC. 2014. Global warming releases microplastic legacy frozen in Arctic Sea ice. Earth’s Future. 2(6):315-320. [cited in PAME 2019].

Oberbeckmann S, Osborn AM, Duhaime MB. 2016. Microbes on a bottle: Substrate, season and geography influence community composition of microbes colonizing marine plastic debris. PLoS One. 11(8):e0159289. 

Oberbeckmann S, Kreikemeyer B, Labrenz M. 2018. Environmental factors support the formation of specific bacterial assemblages on microplastics. Front Microbiol. 8:2709.

Oberbeckmann S, Labrenz M. 2020. Marine microbial assemblages on microplastics: Diversity, adaptation, and role in degradation. Ann Rev Mar Sci. 12:209-232. 

Oberdörster G, Sharp Z, Atudorei V, Gelein EA, Kreyling W, Cox C. 2004. Translocation of inhaled ultrafine particles to the brain. Inhal Toxicol. 16(6-7):437-445.

Oßmann BE, Sarau G, Holtmannspötter H, Pischetsrieder M, Christiansen SH, Dicke W. 2018. Small-sized microplastics and pigmented particles in bottled mineral water. Water Res. 141:307-316.

Ogonowski M, Motiei A, Ininbergs K, Hell E, Gerdes Z, Udekwu KI, Bacsik Z, Gorokhova E. 2018a. Evidence for selective bacterial community structuring on microplastics. Environ Microbiol. 20(8):2796-2808. 

Ogonowski M, Gerdes Z, Gorokhova E. 2018b. What we know and what we think we know about microplastic effects – A critical perspective. Curr Opin Environ Sci Health. 1:41-46.

O’Hara PD, Avery-Gomm S, Wood J, Bowes V, Wilson L, Morgan KH, Boyd WS, Hipfner JM, Desforges JP, Bertram DF, et al. 2019. Seasonal variability in vulnerability for Cassin’s auklets (Ptychoramphus aleuticus) exposed to microplastic pollution in the Canadian Pacific region. Sci Total Environ. 649:50-60. 

Ohtake Y, Kobayashi T, Asabe H, Murakami N. 1995. Biodegradation of low-density polyethylene, polystyrene, polyvinyl chloride, and urea formaldehyde resin buried under soil for over 32 years. J Appl Polym Sci. 56:1789-1796.

Ohtake Y, Kobayashi T, Asabe H, Murakami N, Ono K. 1996. Oxidative degradation and molecular weight change of LDPE buried under bioactive soil for 32-37 years. J Appl Polym Sci. 70:1643-1659.

Özdilek HG, Yalçın-Özdilek Ş, Ozaner FS, Sönmez B. 2006. Impact of accumulated beach litter on Chelonia mydas L. 1758 (Green Turtle) hatchlings of the Samandağ coast, Hatay, Turkey. Fresenius Environ Bull. 15(2):95-103. 

Pacheco A, Martins A, Guilhermino L. 2018. Toxicological interactions induced by chronic exposure to gold nanoparticles and microplastics mixtures in Daphnia magna. Sci Total Environ. 628-629:474-483. 

Page B, McKenzie J, McIntosh R, Baylis A, Morrissey A, Calvert N, Haase T, Berris M, Dowie D, Shaughnessy PD, et al. 2004. Entanglement of Australian sea lions and New Zealand fur seals in lost fishing gear and other marine debris before and after Government and industry attempts to reduce the problem. Mar Pollut Bull. 49(1-2):33-42. 

[PAME] Protection of the Arctic Marine Environment. 2019. Desktop Study on Marine Litter including Microplastics in the Arctic. Akureyri (IS): PAME. 114 p.

Pan Z, Guo H, Chen H, Wang S, Sun X, Zou Q, Zhang Y, Lin H, Cai S, Huang J. 2019. Microplastics in the Northwestern Pacific: abundance, distribution, and characteristics. Sci Total Environ. 650:1913-1922

Panko JM, Chu J, Kreider ML, Unice KM. 2013. Measurement of airborne concentrations of tire and road wear particles in urban and rural areas of France, Japan, and the United States. Atmos Environ. 72:192-199.

Panko JM, Hitchcock KM, Fuller GW, Green D. 2019. Evaluation of tire wear contribution to PM2.5 in urban environments. Atmosphere. 10(2):99.

Panno SV, Kelly WR, Scott J, Zheng W, McNeish RE, Holm N, Hoellein TJ, Baranski EL. 2019. Microplastic contamination in karst groundwater systems. Groundwater. 57:189-196.

Park E-J, Han J-S, Park E-J, Seong E, Lee G-H, Kim D-W, Son H-Y, Han H-Y, Lee B-S. 2020. Repeated-oral dose toxicity of polyethylene microplastics and the possible implications on reproduction and development of the next generation. Toxicol Lett. 324:75-85.

Patterson J, Jeyasanta KI, Sathish N, Booth Am, Edward JKP. 2019. Profiling microplastics in the Indian edible oyster, Magallana bilineata collected from the Tuticorin coast, Gulf of Mannar, Southeastern India. Sci Total Environ. 691:727-735

Pauly JL, Stegmeier SJ, Allaart HA, Cheney RT, Zhang PJ, Mayer AG, Streck RJ. 1998. Inhaled cellulosic and plastic fibers found in human lung tissue. Cancer Epidemiol Biomarkers Prev. 7(5):419-428.

Pedà C, Caccamo L, Fossi MC, Gai F, Andaloro F, Genovese L, Perdichizzi A, Romeo T, Maricchiolo G. 2016. Intestinal alterations in European sea bass Dicentrarchus labrax (Linnaeus, 1758) exposed to microplastics: Preliminary results. Environ Pollut. 212:251-256. 

Peeken I, Primpke S, Beyer B, Gutermann J, Katlein C, Krumpen T, Bergmann M, Hehemann L, Gerdts G. 2018. Arctic sea ice is an important temporal sink and means of transport for microplastic. Nat Commun. 9(1):1505.

Peixoto D, Pinheiro C, Amorim J, Oliva-Teles L, Guilhermino L, Vieira MN. 2019. Microplastic pollution in commercial salt for human consumption: A review. Estuar Coast Shelf Sci. 219:161-168. 

Pellet Watch. 2019. Global monitoring of POPs using beached plastic resin pellets. Tokyo: Laboratory of Organic Geochemistry, Tokyo Univeristy of Agriculture and Technology. [accessed 2019 June 26]. [cited in WHO 2019].

Peng G, Xu P, Zhu B, Bai M, Li D. 2018. Microplastics in freshwater river sediments in Shanghai, China: case study of risk assessment in mega-cities. Environ Pollut. 234:448-456.

Peng J, Wang J, Cai L. 2017. Current understanding of microplastics in the environment: Occurrence, fate, risks, and what we should do. Integr Environ Assess Manag. 13(3):476-482.

Piarulli S, Scapinello S, Comandini P, Magnusson K, Granberg M, Wong JXW, Sciutto G, Prati S, Mazzeo R, Booth AM, et al. 2019. Microplastic in wild populations of the omnivorous crab Carcinus aestuarii: A review and a regional-scale test of extraction methods, including microfibres. Environ Pollut. 251:117-127.

Piehl S, Leibner A, Löder MGJ, Dris R, Bogner C, Laforsch C. 2018. Identification and quantification of macro- and microplastics on an agricultural farmland. Sci Rep. 8(1):17950.

Pieper C, Ventura MA, Martins A, Cunha RT. 2015. Beach debris in the Azores (NE Atlantic): Faial Island as a first case study. Mar Pollut Bull. 101:575-582.

Pierce KE, Harris RJ, Larned LS, Pokras MA. 2004. Obstruction and starvation associated with plastic ingestion in a Northern Gannet Morus bassanus and a Greater Shearwater Puffinus gravis. Mar Ornithol. 32(2):187-189. 

Pigott GH, Ishmael J. 1979. A comparison between in vitro toxicity of PVC powders and their tissue reaction in vivo. Ann Occup Hyg. 22(2):111-126.

Pikuda O, Genbo Xu, E, Berk D, Tufenkji N. 2019. Toxicity assessments of micro- and nanoplastics can be confounded by preservatives in commercial formulations. Environ Sci Technol Lett. 6(1):21-25. 

Pimentel JC, Avila R, Lourenço AG. 1975. Respiratory disease caused by synthetic fibres: a new occupational disease. Thorax. 30(2):204-219.

Pinkerton LE, Yiin JH, Daniels RD, Fent KW. 2016. Mortality among workers exposed to toluene diisocyanate in the US polyurethane foam industry: update and exposure-response analyses. Am J Ind Med. 59(8):630-643. 

Pivokonsky M, Cermakova L, Novotna K, Peer P, Cajthaml T, Janda V. 2018. Occurrence of microplastics in raw and treated drinking water. Sci Total Environ. 643:1644-1651. 

Plot V, Georges JY. 2010. Plastic debris in a nesting leatherback turtle in French Guiana. Chelonian Conserv Bi. 9(2):267-270. 

Poon FE, Provencher JF, Mallory ML, Braune BM, Smith PA. 2017. Levels of ingested debris vary across species in Canadian Arctic seabirds. Mar Pollut Bull. 116(1-2):517-520. 

Porter DW, Castranova V, Robinson VA, Hubbs AF, Mercer RR, Scabilloni J, Goldsmith T, Schwegler-Berry D, Battelli L, Washko R, et al. 1999. Acute inflammatory reaction in rats after intratracheal instillation of material collected from a nylon flocking plant. J Toxicol Environ Health Part A. 57(1):24-45.

Poulain M, Mercier MJ, Brach L, Martignac M, Routaboul C, Perez E, Desjean MC, Halle A. 2019. Small microplastics as a main contributor to plastic mass balance in the North Atlantic Subtropical Gyre. Environ Sci Technol. 53(3):1157-1164.

Prata JC. 2018. Airborne microplastics: Consequences to human health? Environ Pollut. 234:115-126. 

Prata JC, Reis V, Matos JTV, da Costa JP, Duarte AC, Rocha-Santos T. 2019. A new approach for routine quantification of microplastics using Nile Red and automated software (MP-VAT). Sci Total Environ. 690:1277-1283.

Prest EI, Hammes F, van Loosdrecht MCM, Vrouwenvelder JS. 2016. Biological stability of drinking water: Controlling factors, methods, and challenges. Front Microbiol. 7:45.

Provencher JF, Bond AL, Avery-Gomm S, Borrelle SB, Bravo Rebolledo EL, Hammer S, Kühn S, Lavers JL, Mallory ML, Trevail A, et al. 2017. Quantifying ingested debris in marine megafauna: a review and recommendations for standardization. Anal Methods. 9(9):1454-1469.

Provencher JF, Ammendolia J, Rochman CM, Mallory ML. 2018b. Assessing plastic debris in aquatic food webs: what we know and don’t know about uptake and trophic transfer. Environ Rev. p. 1-14.

Provencher JF, Avery-Gomm S, Liboiron M, Braune BM, Macauley JB, Mallory ML, Letcher RJ. 2018c. Are ingested plastics a vector of PCB contamination in northern fulmars from coastal Newfoundland and Labrador? Environ Res. 167:184-190.

Provencher JF, Vermaire JC, Avery-Gomm S, Braune BM, Mallory ML. 2018a. Garbage in guano? Microplastic debris found in faecal precursors of seabirds known to ingest plastics. Sci Total Environ. 644:1477-1484.

Provencher JP, Borrelle SB, Bond AL, Lavers JL, van Franeker JA, Kühn S, Hammer S, Avery-Gomm S, Mallory ML. 2019. Recommended best practices for plastic and litter ingestion studies in marine birds: collection, processing, and reporting. Facets (Ottawa). 4:111-130.

Qi Y, Yang X, Pelaez AM, Huerta Lwanga E, Beriot N, Gertsen H, Garbeva P, Geissen V. 2018. Macro- and micro- plastics in soil-plant system: Effects of plastic mulch film residues on wheat (Triticum aestivum) growth. Sci Total Environ. 645:1048-1056. 

Qiang L, Cheng J. 2019. Exposure to microplastics decreases swimming competence in larval zebrafish (Danio rerio). Ecotoxicol Environ Saf. 176:226-233.

Qiao R, Sheng C, Lu Y, Zhang Y, Ren H, Lemos B. 2019a. Microplastics induce intestinal inflammation, oxidative stress, and disorders of metabolome and microbiome in zebrafish. Sci Total Environ. 662:246-253.

Qiao R, Deng Y, Zhang S, Wolosker MB, Zhu Q, Ren H, Zhang Y. 2019b. Accumulation of different shapes of microplastics initiates intestinal injury and gut microbiota dysbiosis in the gut of zebrafish. Chemosphere. 236:124334.

Rafiee M, Dargahi L, Eslami A, Beirami E, Jahangiri-Rad M, Sabour S, Amereh F. 2018. Neurobehavioral assessment of rats exposed to pristine polystyrene nanoplastics upon oral exposure. Chemosphere. 193:745-753.

Raju S, Carbery M, Kuttykattil A, Senathirajah K, Subashchandrabose SR, Evans G, Thavamani P. 2018. Transport and fate of microplastics in wastewater treatment plants: implications to environmental health. Rev Environ Sci Biotechnol. 17(4):637-653.

Rauert C, Harrad S, Suzuki G, Takigami H, Uchida N, Takata K. 2014. Test chamber and forensic microscopy investigation of the transfer of brominated flame retardants into indoor dust via abrasion of source materials. Sci Total Environ. 493:639-648.

Re V. 2019. Shedding light on the invisible: addressing the potential for groundwater contamination by plastic microfibers. Hydrogeol J. 1-9.

Redondo-Hasselerharm PE, Falahudin D, Peeters ETHM, Koelmans AA. 2018. Microplastic effect thresholds for freshwater benthic macroinvertebrates. Environ Sci Technol. 52(4):2278-2286. 

Rehse S, Kloas W, Zarfl C. 2016. Short-term exposure with high concentrations of pristine microplastic particles leads to immobilisation of Daphnia magna. Chemosphere. 153:91-99. 

Ren X, Tang J, Liu X, Liu Q. 2020. Effects of microplastics on greenhouse gas emissions and the microbial community in fertilized soil. Environ Pollut. 256:113347.

Renzi M, Blašković A. 2018. Litter & microplastics features in table salts from marine origin: Italian versus Croatian brands. Mar Pollut Bull. 135:62-68. 

Renzi M, Guerranti C, Blašković A. 2018. Microplastic contents from maricultured and natural mussels. Mar Pollut Bull Part A. 131:248-251.

Renzi M, Čižmek H, Blašković A. 2019a. Marine litter in sediments related to ecological features in impacted sites and marine protected areas (Croatia). Mar Pollut Bull. 138:25-29.

Renzi M, Grazioli E, Blašković A. 2019b. Effects of different microplastic types and surfactant‑microplastic mixtures under fasting and feeding conditions: A case study on Daphnia magna. Bull Environ Contam Toxicol. 103(3):367-373.

Renzi M, Blašković A, Broccoli A, Bernardi G, Grazioli E, Russo G. 2020. Chemical composition of microplastic in sediments and protected detritivores from different marine habitats (Salina Island). Mar Pollut Bull. 152:110918.

Rezania S, Park J, Md Din MF, Mat Taib S, Talaiekhozani A, Kumar Yadav K, Kamyab H. 2018. Microplastic pollution in different aquatic environments and biota: A review of recent studies. Mar Pollut Bull. 133:191-208.

Richard H, Carpenter EJ, Komada T, Palmer PT, Rochman CM. 2019. Biofilm facilitates metal accumulation onto microplastics in estuarine waters. Sci Total Environ. 683:600-608.

Rillig MC. 2012. Microplastic in terrestrial ecosystems and the soil? Environ Sci Technol. 46(12):6453-6454.

Rillig MC, Ziersch L, Hempel S. 2017. Microplastic transport in soil by earthworms. Sci Rep. 7:1362.

Rist S, Hartmann NB. 2018. Aquatic ecotoxicity of microplastics and nanoplastics: Lessons learned from engineered nanomaterials. In: Wagner M, Lambert S. editors. Freshwater microplastics: emerging environmental contaminants? Cham (CH): Springer. p. 25-49. 

Rist S, Baun A, Almeda R, Hartmann NB. 2019. Ingestion and effects of micro- and nanoplastics in blue mussel (Mytilus edulis) larvae. Mar Pollut Bull. 140:423-430.

Rochman CM. 2015. The complex mixture, fate and toxicity of chemicals associated with plastic debris in the marine environment. In: Bergmann M, Gutow L, Klages M, editors. Marine anthropogenic litter. Cham (CH): Springer. p. 117-140.

Rochman CM, Hoh E, Kurobe T, Teh SJ. 2013. Ingested plastic transfers hazardous chemicals to fish and induces hepatic stress. Sci Rep. 3:3263. 

Rochman CM, Kurobe T, Flores I, Teh SJ. 2014. Early warning signs of endocrine disruption in adult fish from the ingestion of polyethylene with and without sorbed chemical pollutants from the marine environment. Sci Total Environ. 493:656-661. 

Rochman CM, Tahir A, Williams SL, Baxa DV, Lam R, Miller JT, Teh FC, Werorilangi S, Teh SJ. 2015. Anthropogenic debris in seafood: Plastic debris and fibers from textiles in fish and bivalves sold for human consumption. Sci Rep. 5:14340.

Rochman CM, Browne MA, Underwood AJ, van Franeker JA, Thompson RC, Amaral-Zettler LA. 2016. The ecological impacts of marine debris: unraveling the demonstrated evidence from what is perceived. Ecology. 97(2):302-312. 

Rochman CM, Brookson C, Bikker J, Djuric N, Earn A, Bucci K, Athey S, Huntington A, McIlwraith H, Munno K, et al. 2019. Rethinking microplastics as a diverse contaminant suite. Environ Toxicol Chem. 38(4):703-711.

Rosato A, Barone M, Negroni A, Brigidi P, Fava F, Xu P, Candela M, Zanaroli G. 2020. Microbial colonization of different microplastic types and biotransformation of sorbed PCBs by a marine anaerobic bacterial community. Sci Total Environ. 705:135790.

Royer SJ, Ferrón S, Wilson ST, Karl DM. 2018. Production of methane and ethylene from plastic in the environment. PLoS One. 13(8):e0200574.

Rummel CD, Löder MG, Fricke NF, Lang T, Griebeler EM, Janke M, Gerdts G. 2016. Plastic ingestion by pelagic and demersal fish from the North Sea and Baltic Sea. Mar Pollut Bull. 102(1):134-41.

Rummel CD, Jahnke A, Gorokhova E, Kühnel D, Schmitt-Jansen M. 2017. Impacts of biofilm formation on the fate and potential effects of microplastic in the aquatic environment. Environ Sci Technol Lett. 4:258-267.

Ryan PG. 2018. Entanglement of birds in plastics and other synthetic materials. Mar Pollut Bull. 135:159-164.

Ryan PG, Perold V, Osborne A, Moloney CL. 2018. Consistent patterns of debris on South African beaches indicate that industrial pellets and other mesoplastic items mostly derive from local sources. Environ Pollut. 238:1008-1016.

Sadler DE, Brunner FS, Plaistow SJ. 2019. Temperature and clone-dependent effects of microplastics on immunity and life history in Daphnia magna. Environ Pollut. 255:113178.

[SAPEA] Science Advice for Policy by European Academies. 2019. A scientific perspective on microplastics in nature and society. Evidence Review Report No. 4. Berlin (DE): SAPEA. 173 p.

Scherer C, Brennholt N, Reifferscheid G, Wagner M. 2017. Feeding type and development drive the ingestion of microplastics by freshwater invertebrates. Sci Rep. 7:17006.

Scherer C, Weber A, Lambert S, Wagner M. 2018. Interactions of microplastics with freshwater biota. In: Wagner M, Lambert S, editors. 2018. Freshwater microplastics: Emerging environmental contaminants? The Handbook of Environmental Chemistry vol 58. Cham (CH): Springer. p. 153-180.

Schettini E, Stefani L, Vox G. 2014. Interaction between agrochemical contaminants and UV stabilizers for greenhouse EVA plastic films. Appl Eng Agric. 30(2):229-239. 

Scheurer M, Bigalke M. 2018. Microplastics in Swiss floodplain soils. Environ Sci Technol. 52(6):3591-3598.

Schneider T, Burdett G, Martinon L, Brochard P, Guillemin M, Teichert U, Draeger U. 1996. Ubiquitous fiber exposure in selected sampling sites in Europe. Scand J Work Environ Health. 22(4):274-284.

Schrank I, Trotter B, Dummert J, Schölz-Bottcher BM, Löder MJG, Laforsch C. 2019. Effects of microplastic particles and leaching additive on the life history and morphology of Daphnia magna. Environ Pollut. 255:113233.

Schür C, Rist S, Baun A, Mayer P, Hartmann NB, Wagner M. 2019. When fluorescence is not a particle: The tissue translocation of microplastics in Daphnia magna seems an artifact. Environ Toxicol Chem. 38(7):1495-1503.

Schuyler Q, Hardesty BD, Wilcox C, Townsend K. 2014. Global analysis of anthropogenic debris ingestion by sea turtles. Conserv Biol. 28(1):129-139.

Schwabl P, Koppel S, Konigshofer P, Bucsics T, Trauner M, Reiberger T, Liebmann B. 2019. Detection of various microplastics in human stool: A prospective case series. Ann Intern Med. 171(7):453-457. 

Schymanski D, Goldbeck C, Humpf HU, Fürst P. 2018. Analysis of microplastics in water by micro-Raman spectroscopy: release of plastic particles from different packaging into mineral water. Water Res. 129:154-162.

Seiler C, Berendonk TU. 2012. Heavy metal driven co-selection of antibiotic resistance in soil and water bodies impacted by agriculture and aquaculture. Front Microbiol. 3:399.

Selke S, Auras R, Nguyen TA, Castro Aguirre E, Cheruvathur R, Liu Y. 2015. Evaluation of biodegradation-promoting additives for plastics. Environ Sci Technol. 49(6):3769-3777.

Selonen S, Dolar A, Kokalj AJ, Skalar T, Dolcet LP, Hurley R, van Gestel CAM. 2020. Exploring the impacts of plastics in soil - The effects of polyester textile fibers on soil invertebrates. Sci Total Environ. 700:134451.

Seth CK, Shriwastav A. 2018. Contamination of Indian sea salts with microplastics and a potential prevention strategy. Environ Sci Pollut Res Int. 25(30):30122-30131.

Sieber R, Kawecki D, Nowack B. 2020. Dynamic probabilistic material flow analysis of rubber release from tires into the environment. Environ Pollut. 258:113573.

Silva CJM, Silva ALP, Gravato C, Pestana JLT. 2019. Ingestion of small-sized and irregularly shaped polyethylene microplastics affect Chironomus riparius life-history traits. Sci Total Environ. 672:862-868.

Sjollema SB, Redondo-Hasselerharm P, Leslie HA, Kraak MHS, Vethaak AD. 2016. Do plastic particles affect microalgal photosynthesis and growth? Aquat Toxicol. 170:259-261. 

Slootmaekers B, Catarci Carteny C, Belpaire C, Saverwyns S, Fremout W, Blust R, Bervoets L. 2019. Microplastic contamination in gudgeons (Gobio gobio) from Flemish rivers (BE). Environ Pollut. 244:675-684. 

Sobhani Z, Lei Y, Tang Y, Wu L, Zhang X, Naidu R, Megharaj M, Fang C. 2020. Microplastics generated when opening plastic packaging. Sci Rep. 10(1):4841.

Song Y, Cao C, Qiu R, Hu J, Liu M, Lu S, Shi H, Raley-Susman KM, He D. 2019. Uptake and adverse effects of polyethylene terephthalate microplastics fibers on terrestrial snails (Achatina fulica) after soil exposure. Environ Pollut. 250:447-455.

Sorahan T, Pope D. 1993. Mortality and cancer morbidity of production workers in the United Kingdom flexible polyurethane foam industry. Br J Ind Med. 50(6):528-536.

Sørensen L, Rogers E, Altin D, Salaberria I, Booth AM. 2020. Sorption of PAHs to microplastic and their bioavailability and toxicity to marine copepods under co-exposure conditions. Environ Pollut. 258:113844.

Sperling LH. 2006. Introduction to physical polymer science. 4th ed. Hoboken (NJ): John Wiley & Sons, Inc. 845 p.

Stalder T, Top E. 2016. Plasmid transfer in biofilms: A perspective on limitations and opportunities. NPJ Biofilms Microbiomes. 2:16022.

Stamper MA, Spicer CW, Neiffer DL, Mathews KS, Fleming GJ. 2009. Morbidity in a juvenile green sea turtle (Chelonia mydas) due to ocean-borne plastic. J Zoo Wildl Med. 40(1):196-198. 

Stemmer KL, Bingham E, Barkley W. 1975. Pulmonary response to polyurethane dust. Environ Health Perspect. 11:109-113.

Stock V, Böhmert L, Lisicki E, Block R, Cara-Carmona J, Pack LK, Selb R, Lichtenstein D, Voss L, Henderson CJ, et al. 2019. Uptake and effects of orally ingested polystyrene microplastic particles in vitro and in vivo. Arch Toxicol. 93(7):1817-1833.

Stock V, Fahrenson C, Thuenemann A, Dönmez MH, Voss L, Böhmert L, Braeuning A, Lampen A, Sieg H. 2020. Impact of artificial digestion on the sizes and shapes of microplastic particles. Food Chem Toxicol. 135:111010.

Strand J, Feld L, Murphy F, Mackevica A, Hartmann NB. 2018. Analysis of Microplastic Particles in Danish Drinking Water. Aarhus (DK): Aarhus University, Danish Centre for Environment and Energy (DCE). 34 p. Report No. 291.

Su L, Cai H, Kolandhasamy P, Wu C, Rochman CM, Shi H. 2018. Using the Asian clam as an indicator of microplastic pollution in freshwater ecosystems. Environ Pollut. 234:347-355. 

Sun J, Dai X, Wang Q, van Loosdrecht MCM, Ni BJ. 2019. Microplastics in wastewater treatment plants: Detection, occurrence and removal. Water Res. 152:21-37.

Sundet JH, Herzke D, Jenssen M. 2016. Forekomst og kilder i mikroplastikk i sediment, og konsekvenser for bunnlevende fisk og evertebrater på Svalbard. In: Miljøvernfond S, editor. Sluttrapport. Sysselmannen, Norway, p. 1-13. [cited in Hallanger and Gabrielsen 2018].

Sundt P, Shulze PE, Syversen F. 2014. Sources of microplastic pollution to the marine environment [PDF]. Trondheim (NO): Norwegian Environment Agency. [accessed 2019 July 5]. Report No. M-321|2015. 86 p.

Suhrhoff TJ, Scholz-Böttcher BM. 2016. Qualitative impact of salinity, UV radiation and turbulence on leaching of organic plastic additives from four common plastics—a lab experiment. Mar Pollut Bull. 102:84­-94. [cited in WHO 2019].

Sussarellu R, Suquet M, Thomas Y, Lambert C, Fabioux C, Pernet MEJ, Le Goïc N, Quillien V, Mingant C, Epelboin Y, et al. 2016. Oyster reproduction is affected by exposure to polystyrene microplastics. Proc Natl Acad Sci U S A. 113(9):2430-2435.

Talvitie J, Mikola A, Koistinen A, Setälä O. 2017. Solutions to microplastic pollution – Removal of microplastics from wastewater effluent with advanced wastewater treatment technologies. Water Res. 123:401-407.

Tanaka K, Takada H, Yamashita R, Mizukawa K, Fukuwaka M, Watanuki Y. 2013. Accumulation of plastic-derived chemicals in tissues of seabirds ingesting marine plastics. Mar Pollut Bull. 69:219-222.

Tang BL. 2017. Commentary: Tissue accumulation of microplastics in mice and biomarker responses suggest widespread health risks of exposure. Front Environ Sci. 5:63. 

Tang J, Wang X, Yin J, Han Y, Yang J, Lu X, Xie T, Akbar S, Lyu K, Yang Z. 2019. Molecular characterization of thioredoxin reductase in waterflea Daphnia magna and its expression regulation by polystyrene microplastics. Aquat Toxicol. 208:90-97.

Thyssen J, Kimmerle G, Dickhaus S, Emminger E, Mohr U. 1978. Inhalation studies with polyurethane foam dust in relation to respiratory tract carcinogenesis. J Environ Pathol Toxicol. 1(4):501-508.

Toussaint B, Raffael B, Angers-Loustau A, Gilliland D, Kestens V, Petrillo M, Rio-Echevarria IM, Van den Eede G. 2019. Review of micro- and nanoplastic contamination in the food chain. Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 36(5):639-673.

Trestrail C, Walpitagama M, Hedges C, Truskewycz A, Miranda A, Wlodkowic D, Shimeta J, Nugegoda D. 2020. Foaming at the mouth: Ingestion of floral foam microplastics by aquatic animals. Sci Total Environ. 705:135826.

Tunahan Kaya A, Yurtsever M, Çiftçi Bayraktar S. 2018. Ubiquitous exposure to microfiber pollution in the air. 2018. Eur Phys J Plus. 133:488.

Uhl W, Eftekhardadkhah M, Svendsen C. 2018. Mapping microplastic in Norwegian drinking water. Hamar (NO): Norwegian Water. p. 43. Report No. 241/2018. 

Unice KM, Weeber MP, Abramson MM, Reid RCD, van Gils JAG, Markus AA, Vethaak AD, Panko JM. 2019a. Characterizing export of land-based microplastics to the estuary — Part I: Application of integrated geospatial microplastic transport models to assess tire and road wear particles in the Seine watershed. Sci Total Environ. 646:1639-1649.

Unice KM, Weeber MP, Abramson MM, Reid RCD, van Gils JAG, Markus AA, Vethaak AD, Panko JM. 2019b. Characterizing export of land-based microplastics to the estuary – Part II: Sensitivity analysis of an integrated geospatial microplastic transport modeling assessment of tire and road wear particles. Sci Total Environ. 646:1650-1659.

[UNEP] United Nations Environment Programme. 2015. Biodegradable plastics and marine litter: Misconceptions, concerns, and impacts on marine environments. Nairobi (KY): United Nations. 38 p. [accessed 2019 July 5]. 

[UNEP] United Nations Environment Programme. 2016. Marine plastic debris and microplastics: Global lessons and research to inspire action and guide policy change. Nairobi (KY): United Nations. 252 p.

[US EPA] US Environmental Protection Agency. 2002. Health risks from microbial growth and biofilms in drinking water distribution systems [PDF]. Washington (DC): US EPA. [accessed 2019 July].

[US EPA] US Environmental Protection Agency. 2009. Integrated science assessment (ISA) for particulate matter (Final Report, Dec 2009). Washington (DC): US EPA. Report No. EPA/600/R-08/139F.

Van Cauwenberghe L, Janssen CR. 2014. Microplastics in bivalves cultured for human consumption. Environ Pollut. 193:65-70. 

Van Cauwenberghe L, Claessens M, Vandegehuchte MB, Janssen CR. 2015. Microplastics are taken up by mussels (Mytilus edulis) and lugworms (Arenicola marina) living in natural habitats. Environ Pollut. 199:10-17. 

Vassilenko K, Watkins M, Chastain S, Posacka A, Ross PS. 2019. Me, my clothes and the ocean: The role of textiles in microfiber pollution. Science Feature. Vancouver (BC): Ocean Wise. 16 p.

Verschoor AJ. 2015. Towards a definition of microplastics: Considerations for the specification of physico-chemical properties. RIVM Letter report. Bilthoven (NL): National Institute for Public Health and the Environment. 38 p.

Vert M, Doi Y, Hellwich KH, Hess M, Hodge P, Kubisa P, Rinaudo M, Schué F. 2012. Terminology for biorelated polymers and applications (IUPAC recommendations 2012). Pure Appl Chem. 84(2):377-410.

Vidyasakar A, Neelavannan K, Krishnakumar S, Prabaharan G, Sathiyabama Alias Priyanka T, Magesh NS, Godson PS, Srinivasalu S. 2018. Macrodebris and microplastic distribution in the beaches of Rameswaram Coral Island, Gulf of Mannar, Southeast coast of India: A first report. Mar Pollut Bull. 137:610-616.

Vianello A, Jensen RL, Liu L, Vollertsen J. 2019. Simulating human exposure to indoor airborne microplastics using a breathing thermal manikin. Sci Rep. 9:8670. 

Votier SC, Archibald K, Morgan G, Morgan L. 2011. The use of plastic debris as nesting material by a colonial seabird and associated entanglement mortality. Mar Pollut Bull. 62(1):168-172.

Volkheimer G. 1975. Hematogenous dissemination of ingested polyvinyl chloride particles. Ann N Y Acad Sci. 246:164-171.

Von Wintersdorff CJH, Penders J, van Niekerk JM, Mills ND, Majumder S, van Alphen LB, Savelkoul PHM, Wolffs PFG. 2016. Dissemination of antimicrobial resistance in microbial ecosystems through horizontal gene transfer. Front Microbiol. 7:173.

Walkinshaw C, Lindeque PK, Thompson R, Tolhurst T, Cole M. 2020. Microplastics and seafood: lower trophic organisms at highest risk of contamination. Ecotoxicol Environ Saf. 190:110066.

Wang W, Ndungu AW, Li Z, Wang J. 2017. Microplastics pollution in inland freshwaters of China: A case study in urban surface waters of Wuhan, China. Sci Total Environ. 575:1369-1374. [cited in WHO 2019].

Wang W, Yuan W, Chen Y, Wang J. 2018. Microplastics in surface waters of Dongting Lake and Hong Lake, China. Sci Total Environ. 633:539-545.

Wang X, Li C, Liu K, Zhu L, Song Z, Li D. 2020. Atmospheric microplastic over the South China Sea and East Indian Ocean: abundance, distribution and source. J Hazard Mater. 389:121846.

Wang Z, Lin T, Chen W. 2020. Occurrence and removal of microplastics in an advanced drinking water treatment plant (ADWTP). Sci Total Environ. 700:134520.

Warheit DB, Webb TR, Reed KL, Hansen JF, Kennedy GL. 2003. Four-week inhalation toxicity study in rats with nylon respirable fibers: rapid lung clearance. Toxicology. 192(2-3):189-210.

Watts AJR, Urbina MA, Corr S, Lewis C, Galloway TS. 2015. Ingestion of plastic microfibers by the crab Carcinus maenas and its effect on food consumption and energy balance. Environ Sci Technol. 49(24):14597-14604. 

Weber A, Scherer C, Brennholt N, Reifferscheid G, Wagner M. 2018. PET microplastics do not negatively affect the survival, development, metabolism and feeding activity of the freshwater invertebrate Gammarus pulex. Environ Pollut. 234:181-189.

Welden NA, Abylkhani B, Howarth LM. 2018. The effects of trophic transfer and environmental factors on microplastic uptake by plaice, Pleuronectes plastessa, and spider crab, Maja squinado. Environ Pollut. 239:351-358.

Welden NAC, Cowie PR. 2016. Long-term microplastic retention causes reduced body condition in the langoustine, Nephrops norvegicus. Environ Pollut. 218:895-900. 

Weinstein JE, Crocker BK, Gray AD. 2016. From macroplastic to microplastic: Degradation of high-density polyethylene, polypropylene, and polystyrene in a salt marsh habitat. Environ Toxicol Chem. 35(7):1632-1640.

Werner S, Budziak A, van Franeker J, Galgani F, Hanke G, Maes T, Matiddi M, Nilsson P, Ooesterbaan L, Priestland E, et al. 2016. Harm caused by marine litter. Joint Research Centre [JRC] Technical Report. Luxembourg: European Union. 92 p. 

Wesch C, Barthel AK, Braun U, Klein R, Paulus M. 2016. No microplastics in benthic eelpout (Zoarces viviparus): An urgent need for spectroscopic analyses in microplastic detection. Environ Res. 148:36-38.

[WHO] World Health Organization. 2019. Microplastics in drinking-water. Geneva (CH): WHO. 101 p.

Wick P, Malek A, Manser P, Meili D, Maeder-Althaus X, Diener L, Diener PA, Zisch A, Krug HF, von Mandach U. 2010. Barrier capacity of human placenta for nanosized materials. Environ Health Perspect. 118(3):432-436. 

Wieczorek AM, Croot PL, Lombard F, Sheahan JN, Doyle TK. 2019. Microplastic ingestion by gelatinous zooplankton may lower efficiency of the biological pump. Environ Sci Technol. 53:5387-5395.

Wiesheu AC, Anger PM, Baumann T, Niessner R, Ivleva NP. 2016. Raman microspectroscopic analysis of fibers in beverages. Anal Methods. 8:5722-5725.

Windsor FM, Durance I, Horton AA, Thompson RC, Tyler CR, Ormerod SJ. 2019. A catchment-scale perspective of plastic pollution. Glob Chang Biol. 25(4):1207-1221.

Winkler A, Santo N, Ortenzi MA, Bolzoni E, Bacchetta R, Tremolada P. 2019. Does mechanical stress cause microplastic release from plastic water bottles? Water Res. 166:115082.

Wójcik-Fudalewska D, Normant-Saremba M, Anastácio P. 2016. Occurrence of plastic debris in the stomach of the invasive crab Eriocheir sinensis. Mar Pollut Bull. 113(1-2):306-311. 

Woodall LC, Sanchez-Vidal A, Canals M, Paterson GLJ, Coppock R, Sleight V, Calafat A, Rogers AD, Narayanaswamy BE, Thompson RC. 2014. The deep sea is a major sink for microplastic debris. R Soc Open Sci. 1(4):140317.

Wright MS, Loeffler-Peltier GL, Stepanauskas R, McArthur JV. 2006. Bacterial tolerances to metals and antibiotics in metal-contaminated and reference streams. FEMS Microbiol Ecol. 58(2):293-302.

Wright SL, Rowe D, Thompson RC, Galloway TS. 2013. Microplastic ingestion decreases energy reserves in marine worms. Curr Biol. 23(23):R1031-R1033. 

Wright SL, Kelly FJ. 2017. Plastic and human health: A micro issue? Environ Sci Technol. 51(12):6634-6647.

 [WRF] Water Research Foundation. 2017. Distribution system management: Understanding and controlling biofilms [PDF]. Denver (CO): WRF. [updated 2017 July].

Wu F, Wang Y, Leung JYS, Huang W, Zeng J, Tang Y, Chen J, Shi A, Yu X, Xu X, et al. 2020. Accumulation of microplastics in typical commercial aquatic species: A case study at a productive aquaculture site in China. Sci Total Environ. 708:135432.

Wu N, Zhang Y, Zhao Z, He J, Li W, Li J, Xu W, Ma Y, Niu Z. 2020. Colonization characteristics of bacterial communities on microplastics compared with ambient environments (water and sediment) in Haihe Estuary. Sci Total Environ. 708:134876. 

Wu X, Pan J, Li M, Li Y, Bartlam M, Wang Y. 2019. Selective enrichment of bacterial pathogens by microplastic biofilm. Wat Res. 165:114979. 

Xiong X, Tu Y, Chen X, Jiang X, Shi H, Wu C, Elser JJ. 2019. Ingestion and egestion of polyethylene microplastics by goldfish (Carassius auratus): Influence of color and morphological features. Heliyon. 5(12):e03063.

Xu H, Verbeken E, Vanhooren HM, Nemery B, Hoet PHM. 2004. Pulmonary toxicity of polyvinyl chloride particles after a single intratracheal instillation in rats. Time course and comparison with silica. Toxicol Appl Pharmacol. 194(2):111-121.

Yadav MK. 2017. Role of biofilms in environment pollution and control. In: Patra J, Vishnuprasad C, Das G, editors. Microbial biotechnology. Singapore (SG): Springer. p. 377-398.

Yang D, Shi H, Li L, Li J, Jabeen K, Kolandhasamy P. 2015. Microplastic Pollution in Table Salts from China. Environ Sci Technol. 49(22):13622-13627. 

Yoo JW, Doshi N, Mitragotri S. 2011. Adaptive micro and nanoparticles: temporal control over carrier properties to facilitate drug delivery. Adv Drug Deliv Rev. 63(14-15):1247-1256.

Yin L, Chen B, Xia B, Shi X, Qu K. 2018. Polystyrene microplastics alter the behavior, energy reserve and nutritional composition of marine jacopever (Sebastes schlegelii). J Hazard Mater. 360:97-105. 

Yu P, Liu Z, Wu D, Chen M, Lv W, Zhao Y. 2018. Accumulation of polystyrene microplastics in juvenile Eriocheir sinensis and oxidative stress effects in the liver. Aquat Toxicol. 200:28-36. 

Yu F, Yang C, Zhu Z, Bai X, Ma J. 2019. Adsorption behavior of organic pollutants and metals on micro/nanoplastics in the aquatic environment. Sci Total Environ. 694:133643.

Zalasiewicz J, Waters CN, Ivar do Sul JA, Corcoran PL, Barnosky AD, Cearreta A, Edgeworth M, Galuszka A, Jeandel C, Leinfelder R, et al. 2016. The geological cycle of plastics and their use as a stratigraphic indicator of the Anthropocene. Anthropocene. 13:4-17. 

Zbyszewski M, Corcoran PL. 2011. Distribution and degradation of fresh water plastic particles along the beaches of Lake Huron, Canada. Water Air Soil Pollut. 220(1-4):365-372.

Zbyszewski M, Corcoran PL, Hockin A. 2014. Comparison of the distribution and degradation of plastic debris along shorelines of the Great Lakes, North America. J Great Lakes Res. 40:288-299.

Zeri C, Adamopoulou A, Bojanić Varezić D, Fortibuoni T, Kovač Viršek M, Kržan A, Mandic M, Mazziotti C, Palatinus A, Peterlin M, et al. 2018. Floating plastics in Adriatic waters (Mediterranean Sea): From the macro- to the micro-scale. Mar Pollut Bull. 136:341-350.

Zettler ER, Mincer TJ, Amaral-Zettler LA. 2013. Life in the "plastisphere": Microbial communities on plastic marine debris. Environ Sci Technol. 47(13):7137-7146.

Ziajahromi S, Kumar A, Neale PA, Leusch FDL. 2018. Environmentally relevant concentrations of polyethylene microplastics negatively impact the survival, growth and emergence of sediment-dwelling invertebrates. Environ Pollut. 236:425-431. 

Zhang C, Jeong C-B, Lee J-S, Wang D, Wang M. 2019. Transgeneration proteome plasticity in resilience of a marine copepod in response to environmentally relevant concentrations of microplastics. Environ Sci Technol. 53(14):8426-8436. 

Zhang GS, Liu YF. 2018. The distribution of microplastics in soil aggregate fractions in southwestern China. Sci Total Environ. 642:12-20.

Zhang J, Gao D, Li Q, Zhao Y, Li L, Lin H, Bi Q, Zhao Y. 2020a. Biodegradation of polyethylene microplastic particles by the fungus Aspergillus flavus from the guts of wax moth Galleria mellonella. Sci Total Environ. 704:135931.

Zhang J, Wanga L, Kannan K. 2020b. Microplastics in house dust from 12 countries and associated human exposure. Environ Int. 134:105314.

Zhang Y, Gu AZ, Cen T, Li X, He M, Li D, Chen J. 2018. Sub-inhibitory concentrations of heavy metals facilitate the horizontal transfer of plasmid-mediated antibiotic resistance genes in water environment. Environ Pollut. 237:74-82.

Zhang Y, Lu J, Wu J, Wang J, Luo Y. 2020. Potential risks of microplastics combined with superbugs: enrichment of antibiotic resistant bacteria on the surface of microplastics in mariculture system. Ecotoxicol Environ Saf. 187:109852.

Zhang Z, Chen Y. 2020. Effects of microplastics on wastewater and sewage sludge treatment and their removal: A review. Chem Eng J. 382:122955.

Zhou Q, Zhang H, Fu C, Zhou Y, Dai Z, Li Y, Tu C, Luo Y. 2018. The distribution and morphology of microplastics in coastal soils adjacent to the Bohai Sea and Yellow Sea. Geoderma. 322:201-208.

Zubris KAV, Richards BK. 2005. Synthetic fibres as an indicator of land application of sludge. Environ Pollut. 138:201-211.

Zuccarello P, Ferrante M, Cristaldi A, Copat C, Grasso A. Sangregorio D, Fiore M, Oliveri Conti G. 2019. Exposure to microplastics (<10 µm) associated to plastic bottles mineral water consumption: The first quantitative study. Water Res. 157:365-371. 

Zumstein MT, Narayan R, Kohler HE, McNeil K, Sander M. 2019. Dos and do nots when assessing the biodegradation of plastics. Environ Sci Technol. 53(17):9967-9969.

Appendix A: New information published or received through public consultation

Appendix B: Additional occurrence of plastics in the global environment

B-1. Shoreline

Plastic pollution has been detected on shorelines around the world. For example, one study found macroplastics on every beach surveyed on an island in French Polynesia, where plastics accounted for 20% to 100% of all litter items (Connors 2017). 

Chen H et al. (2019) collected marine litter around a tourist city in East China and found that plastic pollutants made up the majority of all floating, benthic, and beached litter. Grocery bags were the most commonly found litter item in all three areas. On shorelines, foams from fishing and aquaculture were found at similar concentrations as grocery bags. The average density of floating microplastics was 36 456 items/km².

In a study by Horn et al. (2019), microplastics were found on all 51 Californian beaches sampled. The average microplastic count was 11.8 items per 100 mL of sediment. Fibres accounted for 95% of the microplastic items. The polymers identified were PP, isotactic PP, atactic PP, polyacrylate, PE, and polyester. 

Ryan et al. (2018) collected litter items sized 2 to 25 mm on South African beaches and reported that plastics comprised 99% of all litter items by number, and that industrial pellets (which form the feedstock of the plastics industry) were the most abundant type of plastic. Typically, pellets enter the environment via accidental spills on land or at sea. Corcoran et al. (2015) found that weather conditions are a factor in industrial pellet accumulation, as is the presence of beached organic material in which they may become entrapped.

Zhou et al. (2018) studied the occurrence of microplastics on beaches adjacent to China’s Bohai Sea and Yellow Sea. Microplastics were both visually identified and analyzed using FTIR, which determined that PE and PP made up the majority of samples. Flakes were the most abundant microplastic, followed by foams, fragments, fibres, pellets, films, and sponges. The abundance of microplastics between sampling sites varied significantly, ranging from 1.3 to 14 712.5 particles/kg dry weight (dw), with an overall average of 740 particles/kg. Similarly, Karthik et al. (2018) studied the occurrence of microplastics on beaches along the southeast coast of India. Microplastic particle concentration along the coast ranged from 2 to 178 particles/m², with a mean of 46.6 particles/m². FTIR analysis identified PE, PP, and PS as the main components of identified plastics. 

Plastics, both macro and micro, are widely found in the Arctic, despite its distance from industrialized and highly populated areas. Plastics have been found in all abiotic environments of the European Arctic, and monitoring of beach litter in the Atlantic Arctic in 2017 revealed that the amount of beach litter varied from a mean of 1 475 items per 100 m in the spring to 195 items per 100 m in the summer months (PAME 2019). 

B-2. Surface water

Plastic pollution is found in fresh and marine surface waters worldwide, and extensive research has been done on the occurrence of microplastics in marine surface waters. A brief summary of selected papers is presented below.

In the United States, Mason et al. (2016) collected surface water samples from Lake Michigan and found an abundance of microplastic particles ranging from about 1 400 to 100 000 particles/km² (mean of 17 267 particles/km²), with 59% of the particles in the size range of 0.355 to 0.999 mm. Microplastic abundance was fairly evenly distributed across the lake surface, despite a seasonal gyre that developed in the southern end of the lake. Fragments dominated, followed by fibres and line, and the most common type of microplastics was PE, followed by PP. A study by Wang et al. (2018) investigated microplastics in freshwater in China. Concentrations in Dongting Lake and Hong Lake ranged from 900 to 4 659 particles/m³, and the concentrations were much higher in the outlet channel between Dongting Lake and the Yangtze River, an area with heavy shipping traffic. The microplastics were mainly PE and PP, and the majority were fibres. Additionally, more than 65% of all microplastics were smaller than 2 mm (Wang et al. 2018).

Surface water samples were collected along the Rhine River in Europe, and microplastics were found in all samples with an average concentration of 892 777 particles/km². A peak concentration of 3.9 million particles/km² was measured in a single sample collected at Rees, in Germany, supporting the finding that higher microplastic concentrations are found near densely populated areas. Most of the microplastics recovered were spheres, followed by fragments (Mani et al. 2015).

Macroplastics have been observed floating on the Arctic Ocean surface, and microplastics have been found in Arctic Ocean surface waters and in the water column. Of the microplastics observed in surface and subsurface waters (to a depth of 6 m), 95% were fibres (Hallanger and Gabrielsen 2018). Plastics may also become entrapped in sea ice, and microplastics levels ranging from 38 to 234 particles/m³ of ice have been measured (Obbard et al. 2014). More recently, Peeken et al. (2018) measured microplastic abundances in Arctic sea ice ranging from 1.1 × 10⁶ particles/m³ to 1.2 × 10⁷ particles/m³, with highly variable concentrations. Most of these microplastics were smaller than 50 μm in size. Of the 17 polymers identified, PE was the most common, with a mean of 48%.

In the Adriatic Sea, Zeri et al. (2018) found significantly higher macroplastic abundance in offshore waters (>4 km) than in inshore waters, but higher abundance of microplastics in nearshore waters (≤4 km) than in offshore waters. The authors collected 22 245 particles of floating microplastics from surface waters, and visually identified 658 floating macroplastics, which accounted for 91.4% of litter items recorded. They found that 29% of the macroplastics was plastic bags, 22% was plastic pieces, 15% was sheets, 13% was fish boxes of expanded PS, 8.8% was cover/packaging, 4.3% was PS pieces, and 1.4% was plastic bottles.

Floating litter collected in Vietnam consisted of a mean of 26% plastics by weight. Of the total plastic mass, 37% was plastic bags, 14% plastic packaging (single-use food containers), and 48% was other plastics, such as plastic bottles, food wrappers, cups, and cutlery (Lahens et al. 2018).

The occurrence and aggregation potential of microplastics in the Mediterranean Sea has been reported by several researchers. For example, de Haan et al. (2019) collected surface water samples using 335 µm mesh nets, which yielded 2 489 plastic particles. Microplastics made up 94.6% of plastic abundance and 55% of plastics by weight, averaging 0.10 items/m². The three most abundant polymers were LDPE and HDPE (54.5%), PP (16.5%) and PS (9.7%) (de Haan et al. 2019). 

Bordós et al. (2019) examined microplastic occurrence in Hungary. Given the use of a 2 mm pre-filter during sampling, microplastics between 2 mm and 5 mm were not sampled. Suspected plastic particles were visually identified and analyzed under a FTIR microscope and six polymer types were identified: PE, PP, PS, PTFE, polyacrylate, and polyester. Of the 13 water samples taken, 12 contained microplastics ranging from 3.52 to 32.1 particles/m³ with an average of 13.8 particles/m³. All water entering sampling locations (i.e., influents) had higher microplastic concentrations than the water leaving that sampling location (i.e., effluents).

Pan et al. (2019) reported microplastics in surface waters across the northwestern Pacific Ocean. The concentration of particles collected from 18 stations varied significantly, ranging from 6.4 x 10² to 4.2 x 10⁴ items/km². Microplastics were analyzed by Micro-Raman spectroscopy, yielding a distribution of 57.8% PE, 36.0% PP, and 3.4% nylon.

Poulain et al. (2019) investigated the concentration of microplastics in the North Atlantic Subtropical Gyre. Microplastics were categorized as small microplastics (SMPs, 0.025 to 1 mm) and large microplastics (LPMs, 1 to 5 mm). SMPs were collected by a 25 μm mesh net, and LMPs by a 300 μm mesh net. The authors accounted for the decreased buoyancy of SMPs compared to LMPs and applied a correction factor for the increased susceptibility of microplastics to wind-driven vertical transport. The concentrations of LMPs and SMPs corrected for vertical transport are 50 to 1000 g/km² and 5 to 14 000 g/km², respectively. 

Eriksen et al. (2014) conducted 680 net tows of global surface water and found plastics in 92.3% of the tows. Visual surveys in the South Pacific, North Pacific, South Atlantic, Indian Ocean, and water around Australia also indicated that foamed PS items were the most frequently observed macroplastics. The authors estimated that there are 5.25 trillion particles of plastic floating at sea, totalling 268 940 tonnes. Their results indicate that plastic pollution has spread throughout the world's oceans and that plastics accumulate in subtropical gyres. There is an area with accumulation of buoyant plastics in the North Pacific Subtropical Gyre that is commonly referred to as the Great Pacific Garbage Patch (Eriksen et al. 2014). Lebreton et al. (2018) predicted that a 1.6 million km² zone of the Great Pacific Garbage Patch contains 1.8 trillion pieces of plastics and weighs 79 000 tonnes. The average plastic mass concentration measured inside the Great Pacific Garbage Patch has shown exponential increase over the last decades, from 0.4 kg/km² (n = 20) in the 1970s to 1.23 kg/km² (n = 288) in 2015 (Lebreton et al. 2018).

B-3. Benthic zone

Plastic pollution has also been detected in marine sediments around the world and is typically dominated by microplastics. Dai et al. (2018) reported the occurrence of microplastics in the surface water, water column and sediment of the Bohai Sea of the Pacific Ocean. Microplastics were detected in all 20 surface water samples, ranging from 0.4 to 5.2 particles/L, with an average of 2.2 particles/L. The average concentration of microplastics in the water column ranged from 1.6 to 6.9 particles/L. There was no clear trend in microplastic accumulation at any specific depth along the water column, and the abundance in sediments was inconsistent with the water column. The surface sediment concentration ranged from 31.1 to 256.3 particles/kg. Fibres dominated the type of microplastics found in both water and sediment, followed by fragments. µ-FTIR analysis identified that the polymer with the highest density in surface waters was PS, whereas PET and PVC were found at highest densities in deeper water.

In Argentina, an average of 25 macroplastic items/m² and 704 microplastic fragments/m² were collected from sediment. The macroplastic pollutants were categorized into 24 types, and the most dominant types were food wrappers (PP and PS), bags (HDPE and LDPE), bottles (PET), and disposable Styrofoam food containers (PS) (Blettler et al. 2017). In a study in the United Kingdom, the main types of macroplastic pollution found in sediment were packaging, fishing and shipping waste (Browne et al. 2010). Macroplastics and microplastics were found in sediments from a marine protected area in Italy, ranging from a mean of 11.9 to 46.4 pieces and 151.0 to 678.7 pieces per kg dw of sediment, respectively. Greater than 85% of the microplastics were fibres (Fastelli et al. 2016). Bordós et al. (2019) sampled sediment in Hungary and found that 9 of the 12 sediment samples contained microplastics ranging from 0.46 to 1.62 particles/kg, with an average of 0.81 particles/kg. The most dominant polymer was PP. Marine litter in Croatia ranged from 3.4 items/kg dw to 528 items/kg dw, with macroplastics making up 1.3% to 11.3% of samples. Like in Italy, fibres were the most abundant type of microplastic found in Croatia, ranging from 39.9% to 90.1% of the total number of plastic items (Renzi et al. 2019a; Blašković et al. 2017). Blašković et al. (2017) found no correlation between the extent and pattern of plastic contamination and sediment grain size or sampling depth. In Svalbard in the Arctic, fibres were once again found to be the dominant microplastic in sediment at depths of 40 to 79 m, where they were sampled at a density of 9.2 fibres/kg (Sundet et al. 2016).

Vidyasakar et al. (2018) conducted the first study on the distribution and characteristics of plastic pollutants in marine sediment on Rameswaram Island, along the southeast coast of India. PP was the most abundant polymer type, followed by PE, PS, nylon, and PVC. Irregularly shaped plastics were most plentiful at 69.2%, followed by fibres at 17.9% and pellet-shaped plastics at 12.9%.

Microplastics have been found in large quantities in river sediment in Shanghai (Peng et al. 2018), at concentrations ranging from 5.3 particles/100 g dw to 160 particles/100 g dw. The average concentration across all sites was 80.2 particles/100 g dw. Residential areas showed the highest level of microplastic concentration, followed by parks, rural areas, and tourist areas. Spheres constituted the majority of microplastics at 88.98%, followed by fibres (7.55%) and fragments (3.47%). The two most dominant polymer types identified by μ-FTIR were PP and polyesters (Peng et al. 2018).

García-Rivera et al. (2018) derived data from the MEDITS (International Bottom Trawl Survey in the Mediterranean) program surveys and found that, over 11 years, 2197.8 kg of marine litter was collected from the Spanish Mediterranean seafloor (collected five stratum levels at depths from 0 to 800 m) and was comprised of 29.3% plastics by weight. They reported that the amount of marine litter generally remained stable over the survey period. Deep sea litter in the Arctic reportedly increased from 346 items/km² in 2004 to 8082 items/km² in 2014, with plastics accounting for 47% of litter (PAME 2019). A plastic bag was found in the Mariana Trench at a depth of 10 898 m (Chiba et al. 2018).

Appendix C: Additional information on occurrence of microplastics in food

^a Mussels are the most frequently investigated species of mollusc. Similar concentrations of microplastics have been reported in clams, oysters, scallops, and snails.
^b Microplastics were only observed in the digestive tract, head, and gills of the whole shrimp and not in the abdominal muscle tissue of peeled shrimp.
^c Microplastic concentrations in salt varied considerably depending on the origin and type of salt.

Appendix D: Additional information on ecotoxicological studies

^a Median lethal time
^b Wet weight
^c Hours post fertilization
^d Quantitative polymerase chain reaction
^e Glutathione
^f Glutathione peroxidase
^g Superoxide dismutase
^h Glutathione-S-transferase
^I Mitogen-activated protein kinase
^j Extracellular signal-regulated kinase
^k Protein kinase B
^l Mitogen-activated protein kinase (MAPK) kinase

^a Reactive oxygen species
^b Glutathione reductase

^a Dry weight
^b Median effective concentration
^c 10% effect concentration

Appendix E: Additional information on toxicological studies

^a No observed effect level
^b Lactate dehydrogenase
^c Carboxymethylcellulose

^a Geometric mean
^b Lowest observed effect concentration
^c Lowest observed effect concentration, adjusted for continuous exposure
^d No observed effect concentration
^e No observed effect concentration, adjusted for continuous exposure
^f Mass median aerodynamic diameter
^g Lowest observed effect level
^h Gestational day