Abstract
Microplastics are one of the emerging contaminants that have received attention in recent decades due to their adverse effects on human health and the environment. Though microplastics are primarily found in abundance in oceans, freshwater sources and drinking water are not unaffected. Nevertheless, it is not only the microplastics that are harmful; rather their ability to transport contaminants is another serious issue of concern. The contaminant transport ability is affected by various environmental and physico-chemical parameters of microplastics. Lack of effective and targeted water treatment technologies have led intake of microplastics by humans resulting in a variety of health issues. Even though a few regulatory attempts have been made in the direction of curtailing the production and use of microplastics, there is still a long way to go. This paper focuses on various aspects of microplastics’ presence in drinking water, focusing on their contaminant transport ability, human health risks, removal technologies, and the global scenario of concern.
HIGHLIGHTS
Global production of plastics has increased from 1.5 million tons in 1950 to 359 million tons in 2018.
Microplastics act as the carrier of a variety of organic/inorganic contaminants and pathogens in drinking water.
Contaminant transport ability is influenced by environmental and physico-chemical properties of microplastics.
Microplastics in drinking water might trigger a variety of adverse health impacts.
Graphical Abstract
INTRODUCTION
Plastics are an important material in today's world considering their remarkable physical and chemical properties, inertness, wide-scale usability, and reusability. Realizing the potential of plastics, its production has increased globally from 1.5 million tons in 1950 to 359 million tons in 2018 (Garside 2019; Plastics Europe 2019) (Figure 1). Despite having numerous benefits, plastics are associated with myriad environmental and human health issues due to their extremely slow biodegradation in nature and release of poisonous chemicals upon burning (Thornton et al. 1996; Sindiku et al. 2015; Ni et al. 2016; Rajmohan et al. 2019). In many of the countries a significant portion of plastic waste is discarded as such, instead of being recycled or incinerated (Geyer et al. 2017; Plastics Europe 2019) (Figure 1). The discarded plastic waste accumulated over land sources, agricultural land, stagnant water bodies, and wastewater effluent/sludge, finds its way into the rivers. It has been estimated that approximately 70–80% of plastic waste from land based sources, which is approximately 1.15–2.41 million tons, is carried away by the rivers, and ultimately ends up in the oceans (Horton et al. 2017; Lebreton et al. 2017). The major contributors are the Asian rivers amounting to approximately 67% of the global total (Lebreton et al. 2017). Breakdown of this plastic waste over time through the action of waves and winds results in the formation of microplastics. Microplastic is a relatively new term and an emerging contaminant, which is defined as synthetic plastic polymer having an upper size limit of 5 mm (Arthur et al. 2009; Thompson et al. 2009). While dealing with discarded plastic waste is still a significant global public health challenge, discovering the role of microplastics as an emerging pollutant of concern is another issue for the environmental health stakeholders.
Based on their origin, microplastics can be categorized into primary and secondary microplastics. Primary microplastics are intentionally manufactured in sizes <5 mm to be used in various applications such as cosmetics, clothing and other textiles, fishing nets, etc. (Mai et al. 2018). However, secondary microplastics originate from the breakdown of discarded plastic waste by solar radiation, mechanical degradation, microbial action, etc. (Rodrigues et al. 2018; Wagner & Lambert 2018). These microplastic particles may be of various shapes such as fragments, pellets, beads, and fibres. The occurrence of microplastics has been reported from oceans, sediments, freshwater, wastewater, tap water, bottled water, air, food products, aquatic organisms, etc. (WHO 2019). The prevalent units used to express microplastics’ abundance in water, sediment, and biota are particles/m3 (or particles/L), particles/m2, and particles/individual, respectively (Mai et al. 2018).
The identification and quantification of microplastics have been carried out using various techniques such as microscopy, spectroscopy, and chromatography. The simplest and the most common technique to identify and quantify microplastic particles is optical microscopy, where quantification is achieved through manual counting. Though this technique is simple, it poses limitations in terms of misidentification and reduced accuracy (underestimation (Loder et al. 2015)/overestimation (Lenz et al. 2015)). However, application of electron microscopic techniques, such as scanning electron microscopy, may overcome this limitation to some extent (Eriksen et al. 2013). Application of spectroscopic analysis is more pertinent considering that spectroscopic techniques can identify the chemical composition of microplastics and hence can differentiate among the varieties of microplastic particles. Fourier transform infrared spectroscopy (Kosuth et al. 2018; Schymanski et al. 2018; Mintenig et al. 2019) and Raman spectroscopy (Oβmann et al. 2018; Pivokonsky et al. 2018) are utilized for this purpose. The combination of microscopic and spectroscopic techniques further enhances the output (Pivokonsky et al. 2018). Apart from these, researchers have used some new techniques as well for the identification of microplastics, such as pyrolysis gas chromatography mass spectroscopic (GC-MS) techniques (Fries et al. 2013; Peters et al. 2018; Fischer & Scholz-Bottcher 2019; Funck et al. 2020). In contrast to microscopic and spectroscopic techniques, the pyrolysis GC-MS technique is a destructive one, where a sample needs to be pyrolyzed to get it identified. Moreover, this technique fails to provide information regarding the number, type, and morphology of the particles (Hanvey et al. 2017). Pressurized fluid extraction is another technique for the quantification of microplastic particles where semi-volatile organics are obtained from solid plastic materials at sub-critical temperature and pressure (Fuller & Gautam 2016). A detailed review on identification and quantification techniques of microplastics is provided by Hanvey et al. (2017).
The microplastics are of immense environmental and public health concern because of their inherent physico-chemical properties and ubiquitous presence as well as persistence (Rezania et al. 2018; Xu et al. 2018a). Moreover, microplastics also leach out various persistent organic pollutants, which are of serious health concern (Silva et al. 2018). Microplastics may pose health impacts in three ways: (1) the particles themselves, as these are small and can be ingested/inhaled easily; (2) various chemicals which are present in microplastics such as sorbed moieties, additives, etc.; and (3) microorganisms, which get attached to and colonize over the microplastics and result in biofilm formation (WHO 2019). Thus, microplastics act as a carrier of various hazardous moieties/organisms and affect biotic species including the human beings. Their contaminant transport ability is governed by various environmental and physico-chemical properties. For example, acidic pH helps in adsorption of a variety of the contaminants (Holmes et al. 2014; Li et al. 2019; Wang et al. 2020a), while the presence of dissolved organic matter reduces the adsorption (Wu et al. 2016; Xu et al. 2018b). This happens because pH and dissolved organic matter affect the basic properties of chemicals (contaminants) such as pKa and hydrophilicity. Similarly, small particle size and low density of microplastics promote the sorption of contaminants and vice-versa (Mato et al. 2001, 2002; Teuten et al. 2007; Liu et al. 2018a; Wang et al. 2018; Li et al. 2019).
The occurrence of microplastics in marine environment is widely recognized and documented (Barboza & Gimenez 2015; Galloway & Lewis 2016; Wright & Kelly 2017; Carbery et al. 2018; Karthik et al. 2018; Rezania et al. 2018; Smith et al. 2018; Wang et al. 2018; Ashwini & Varghese 2020; Robin et al. 2020). However, there are very few comprehensive studies/reports about the occurrence of microplastics in drinking water (Kosuth et al. 2018; Schymanski et al. 2018; Eerkes-Medrano et al. 2019; Koelmans et al. 2019; Novotna et al. 2019). Considering the human health aspects, drinking water is one of the most important sources through which microplastics can be ingested. Therefore, it is necessary to understand various environmental and physico-chemical factors that might affect the microplastics’ transport and possible health impacts of microplastics found in drinking water. In this paper, occurrence of microplastic particles in drinking water has been reviewed. Aspects covered in this context are – the origin of microplastics, environmental and physico-chemical factors associated with microplastics’ occurrence and contaminating ability, human health impacts, possible removal methods, and actions taken for curtailing the production and use of the same.
MICROPLASTICS: ORIGIN AND PATHWAYS TO REACH UP TO DRINKING WATER
Microplastics (both primary and secondary) pollute drinking water sources primarily through discharge of sewage/wastewater treatment plant effluent and surface run-off (Figure 2). There are large numbers of industries that use (primary) microplastics for various applications, such as medicines, cosmetics, etc. After their use, these primary microplastics get washed off and become a part of the domestic wastewater (Singh et al. 2021). As the sewage/wastewater treatment plants are not equipped for the complete removal of microplastics, the effluent released from these plants contains substantial quantity of microplastics (Amrutha & Warrier 2020). Upon mixing of this effluent with the freshwater sources, microplastics become part of the fresh/drinking water supply chain (Magnusson & Noren 2014; Novotna et al. 2019; Okoffo et al. 2019). For example, increase in the concentration of microplastics in the Chicago River has been reported to be due to the local wastewater treatment plants’ effluent (McCormick et al. 2014). It is also important to note that many components of water treatment plants and water distribution system are usually made up of plastic materials, such as high density polyethylene, polyvinyl chloride, polypropylene, etc. (Mintenig et al. 2019) and hence, these further contribute towards microplastic generation in the water they carry.
Treated bottled water is also reported to contain microplastics (Mason et al. 2018; Oβmann et al. 2018; Pivokonsky et al. 2018; Schymanski et al. 2018). Mason et al. have reported that approximately 10.4 microplastic particles have been found in a litre of bottled drinking water, with size of more than 100 μm (Mason et al. 2018). Nevertheless, the smallest microplastic particle reported in case of drinking water is 1 μm (WHO 2019). Evidence suggests that it is the bottling process and/or packaging of the plastic bottles/caps that largely contributes to the generation of microplastics. It has been found that water from the same source shows more microplastics when it is packaged in plastic bottles, as compared to one packaged in glass bottles (Mason et al. 2018; Schymanski et al. 2018). Polypropylene is the most common type of microplastic found in bottled water (Mason et al. 2018). However, another study revealed that a mix of polyester and polyethyleneterephthalate is most commonly found in single-use plastic bottles (Schymanski et al. 2018). By and large, microplastics in drinking water are the outcome of improperly disposed plastic materials and plastic packaging. A summary of the occurrence of microplastic particles in drinking water is presented in Table 1.
Type of water . | Identification method . | Size range (μm) . | Concentration (particles/L) . | Morphology . | Composition . | References . |
---|---|---|---|---|---|---|
Ground drinking water sources | FTIR spectroscopya | 50–150 | 0–7 | Fragments | Polyethylene, polyamide, polyester, polyvinylchloride | Mintenig et al. (2019) |
Pyrolysis – GC MSa | – | 6.4 | Fibres | Polyethylene | Panno et al. (2019) | |
Treated water from water treatment plants | FTIR spectroscopy & micro-Raman imaging microscopy | 1–10 | 338±76 to 628±28 | Fragments, fibres | Polyethylene terephthalate, Polypropylene, Polyethylene | Pivokonsky et al. (2018) |
Tap water | FTIR spectroscopy | 100–5,000 | 0–61 | Fibres | – | Kosuth et al. (2018) |
Plastic bottled water | FTIR spectroscopy | 6.5 – >100 | 0 to >10,000 | Fragments, fibres | Polypropylene | Mason et al. (2018) |
– | 3.57 | Fibres | – | Kosuth et al. (2018) | ||
Micro – FTIR spectroscopy | 5–20 | 118±88 (returnable bottles), 14±14 (single-use bottles) | Fragments | Polyethylene terephthalate, Polypropylene | Schymanski et al. (2018) | |
Micro – Raman spectroscopy | < 5 | 4,890±5,430 (returnable bottles) 2,649±2,857 (single-use bottles) | – | Polyethylene | Oβmann et al. (2018) | |
Glass bottled water | FTIR spectroscopy | 6.5 – >100 | 1,410 14.8 (>100 μm) 1,396 (6.5–100 μm) | Fragment, fibre, pellet, film, foam | – | Mason et al. (2018) |
Micro – FTIR spectroscopy | >100 | 50±52 | – | Polyamide, Polyethylene, Polypropylene | Schymanski et al. (2018) | |
Micro – Raman spectroscopy | <5 | 6,292±10,521 | – | Polyethylene, Styrene – butadiene – copolymer | Oβmann et al. (2018) | |
Cardboard bottled water | Micro – FTIR spectroscopy | >100 | 11±8 | Fibres | Cellulose, Polyethylene, Polypropylene | Schymanski et al. (2018) |
Type of water . | Identification method . | Size range (μm) . | Concentration (particles/L) . | Morphology . | Composition . | References . |
---|---|---|---|---|---|---|
Ground drinking water sources | FTIR spectroscopya | 50–150 | 0–7 | Fragments | Polyethylene, polyamide, polyester, polyvinylchloride | Mintenig et al. (2019) |
Pyrolysis – GC MSa | – | 6.4 | Fibres | Polyethylene | Panno et al. (2019) | |
Treated water from water treatment plants | FTIR spectroscopy & micro-Raman imaging microscopy | 1–10 | 338±76 to 628±28 | Fragments, fibres | Polyethylene terephthalate, Polypropylene, Polyethylene | Pivokonsky et al. (2018) |
Tap water | FTIR spectroscopy | 100–5,000 | 0–61 | Fibres | – | Kosuth et al. (2018) |
Plastic bottled water | FTIR spectroscopy | 6.5 – >100 | 0 to >10,000 | Fragments, fibres | Polypropylene | Mason et al. (2018) |
– | 3.57 | Fibres | – | Kosuth et al. (2018) | ||
Micro – FTIR spectroscopy | 5–20 | 118±88 (returnable bottles), 14±14 (single-use bottles) | Fragments | Polyethylene terephthalate, Polypropylene | Schymanski et al. (2018) | |
Micro – Raman spectroscopy | < 5 | 4,890±5,430 (returnable bottles) 2,649±2,857 (single-use bottles) | – | Polyethylene | Oβmann et al. (2018) | |
Glass bottled water | FTIR spectroscopy | 6.5 – >100 | 1,410 14.8 (>100 μm) 1,396 (6.5–100 μm) | Fragment, fibre, pellet, film, foam | – | Mason et al. (2018) |
Micro – FTIR spectroscopy | >100 | 50±52 | – | Polyamide, Polyethylene, Polypropylene | Schymanski et al. (2018) | |
Micro – Raman spectroscopy | <5 | 6,292±10,521 | – | Polyethylene, Styrene – butadiene – copolymer | Oβmann et al. (2018) | |
Cardboard bottled water | Micro – FTIR spectroscopy | >100 | 11±8 | Fibres | Cellulose, Polyethylene, Polypropylene | Schymanski et al. (2018) |
aFTIR: Fourier Transform Infra-Red (FTIR) Spectroscopy; GC MS: Gas Chromatography Mass Spectrometry.
Though a significant portion of microplastics in drinking water is contributed by the primary microplastics (∼ 15–31%), the proportion contributed by the secondary microplastics is huge, which amounts to approximately 69–81% (EU 2018). These secondary microplastics originate through the breakdown of macroplastics in landfills and open dumping sites (Rillig 2012; Duis & Coors 2016; Liu et al. 2018b; Rodriguez-Seijo & Pereira 2018). The leachate generated from these sites contains significant amounts of microplastics which seep along with the water and contaminate the freshwater as well as groundwater sources. This is the reason why not only the drinking water obtained from surface water sources is contaminated, but also the water extracted from the ground is found to contain microplastics, though in lesser quantities (Mintenig et al. 2019).
Breakdown of macroplastics to (secondary) microplastics is largely influenced by the environmental factors, such as sunlight, temperature, air/oxygen, etc. The impact of these factors on the breakdown process indicates that the process varies from place to place, as different locations have different environmental features (Weinstein et al. 2016). It has been seen that polyethylene materials degrade faster in air, as compared to seawater, because of the difference in temperature, oxygen content, and sunlight exposure (Pegram & Andrady 1989; Andrady et al. 1993). Moreover, weathering is slow in submerged habitats because of the fast reduction of UV-B radiation, biofilm formation, low temperature, and oxygen content (Andrady 2011).
MICROPLASTICS: CARRIER OF CONTAMINANTS
Role of microplastics as carrier of contaminants
The microplastics are not only harmful per se, but act as source and sink for many other contaminants as well. Microplastics themselves act as source of various chemicals, additives, and pigments, which are added during the manufacturing process of the plastics. Further, microplastics also adsorb a variety of inorganic and organic contaminants such as persistent organic pollutants, metals/metalloids, pesticides, pharmaceuticals, microorganisms, etc. and hence also act as a sink. Several authors have reported that microplastics act as an important carrier for transportation of heavy metals, particulate matter, persistent organic pollutants, etc. (Brennecke et al. 2016; Kwon et al. 2017; Yu et al. 2019; Barbosa et al. 2020). Moreover, the sorption depends on various factors such as, physicochemical properties of the polymer, surrounding environment, atmospheric temperature, humidity, salinity, weathering, ageing processes, etc. As microplastics get weathered with time due to the effect of various environmental factors, their ability to transport contaminants also gets affected (Hartmann et al. 2017) which further enhances threat for the drinking water sources. Additionally, plastic debris gain more surface area upon weathering, thus generating oxygen groups that collectively affect their polarity, roughness, charge, and porosity; thereby enhancing the adsorption capacity (Fotopoulou & Karapanagioti 2012). Surface charge, surface area, functional groups, and acid-base characteristics also influence the sorption of contaminants on the microplastics’ surface (Fred-Ahmadu et al. 2020).
Colonization of microorganisms over microplastics present in the aquatic system is another important aspect that has been looked into (Harrison et al. 2014). Microplastic surfaces are known to promote the survival and growth of a wide variety of microorganisms (Wu et al. 2019). Various antibiotic-resistant genes as well as human pathogens have been detected on the surface of microplastic particles (Wu et al. 2019). The first report of microplastic colonization by microbes appeared in the 1970 s, where white plastic pellets were found to be associated with diatoms and hydroids (Carpenter & Smith 1972; Carpenter et al. 1972). Virsek et al. have reported the presence of the bacterial fish pathogen Aeromonas salmonicida (syn Haemophilus piscium) on the surface of microplastics (Virsek et al. 2017). The reason behind the attachment of microbes on the microplastics is that any solid surface present in an aquatic environment is prone to absorb a variety of nutrients that attract the microbes (Oberbeckmann et al. 2015). An excess of nutrients over the microplastic particles acts as hotspots and competition may also develop among the microbes to obtain it (Hall-Stoodley et al. 2004; Salta et al. 2013). The microbe-laden microplastics, when ingested by humans, have the potential to cause various health-related issues such as endocrine disruption, cytotoxicity, etc.
The biofilm formation is also found responsible for the adsorption of metals over the microplastics’ surface. It has been shown that all the metals, irrespective of space and time, do get adsorbed on the microplastics (Johansen et al. 2018). Later it was suggested that this accumulation might be controlled by the biofilm formation, and the distribution of biofilm is similar among the variety of plastics. Therefore, it allows all types of metals to get adsorbed on most of the plastics (Johansen et al. 2018).
Effect of environmental parameters on the contaminant transport capacity of microplastics
Many environmental parameters affect the contaminant transport capacity of the microplastics, such as, pH, temperature, salinity, organic matter, etc. (Figure 3). pH is one of the important factors to consider as it has been reported that high pH results in decreased sorption of some antibiotics and surfactants on the surface of metals, while decreasing pH has the opposite effect (Holmes et al. 2012, 2014; Wang et al. 2015; Guo et al. 2018). Increase in pH results in a high concentration of hydroxide (OH-) ions. Interaction of OH- ion with the various ionic forms of the contaminant decides the sorption behaviour. The ionic strength of the medium is another important parameter that represents the charge associated with it (Zhang et al. 2018; Hu et al. 2020). This property might increase or decrease the sorption depending upon the ionic strength of the medium and type of contaminant involved, as well as synergistic and/or antagonistic interaction with other parameters (Brewer et al. 2020). The mechanism involved here is the shrinkage of the plastic polymer upon increasing the ionic strength, which causes a reduction in the pore size as well as the number of adsorption sites, thereby reducing the adsorption (Xu et al. 2008; Dong et al.2020a). Solar radiation also results in the breaking of bonds in the polymers, which increases the surface area and pore size of the microplastics, thereby allowing more adsorption of organic moieties over microplastics (Kalogerakis et al. 2017).
Temperature and salinity are other environmental parameters that affect the sorption of contaminants on the microplastics. The vander Waals force among the molecules is the deciding factor, as it decreases upon increasing the temperature owing to faster mobility and solubility of molecules at higher temperatures (Gusso & Burnham 2016) which results in enhanced sorption (Table 2). Contrary to this, it has also been seen that after a certain point, high temperature also reduces surface tension, which ultimately results in low sorption of contaminants (Liu et al. 2018a). As far as salinity is concerned, it might not affect the freshwater sources directly; however, it affects the microplastics’ contaminant carrying ability in marine environment (Hu et al. 2017; Liu et al. 2018a, 2019). Thus, microplastics that pass on from marine environment to surface/ground water resources through mixing, sea-water intrusion, and/or through biotic species, are of concern. It has been reported that with increasing salinity, lubricating oil gets adsorbed more on the polyethylene and polystyrene microplastics because of the ease of the outer-sphere surface complexation provided by the salts (Hu et al. 2017). Enhancement in the sorption capacity of contaminants over microplastics upon increasing the salinity implies that microplastics occurring in the marine environment are more prone to carry the contaminants (Velzeboer et al. 2014; Wang et al. 2015). However, this effect varies in the case of metals. The adsorption capacity for a few metals decreases with an increase in salt content, while it increases for others (Table 2). It has been postulated that competition for the sorption sites on the microplastic pellets is the determining factor for sorption (Holmes et al. 2014; Liu et al. 2018a). Influence of salinity also results in alteration in the agglomeration behaviour of microplastics which further influences the size and area related properties (Velzeboer et al. 2014).
Type of microplastic . | Type of contaminant sorbed . | Change in parameter . | Impact on contaminant sorption on MPs . | References . |
---|---|---|---|---|
pH | ||||
Polystyrene | Metalloid (Arsenic) | Increase (pH 3→7) | Decreased | Dong et al. (2020a) |
Polyethylene | Pesticides (Carbendazim, Dipterex) | Increase (pH 2→6) | Decreased | Wang et al. (2020a) |
Pesticides (Malathion) | Increase (pH 2→3) | Increased | ||
Pesticides (Diflubenzuron, Difenoconazole) | Increase (pH 2→6) | Increased | ||
Polytetrafluoroethylene | Metalloid (Arsenic) | Increase (pH 3→7) | Decreased | Dong et al. (2019) |
Polystyrene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Increase (pH 3→6) | Increased | Li et al. (2019) |
Increase (pH 6→12) | Decreased | |||
Polystyrene, Polyvinylchloride | Antibiotic (Tylosin) | Increase (pH 3→7) | Decreased | Guo et al. (2018) |
Polyethylene, Polypropylene, Polystyrene | Antibiotic (Tetracycline) | Increase (pH 2→6) | Increased | Xu et al. (2018b) |
Polyethylene | Antibiotic (Sulfamethoxazole) | Increase (pH 2→12) | Increased (slightly) | Xu et al. (2018c) |
Polyethylene, Polystyrene | Lubrication oil | Increase (pH 1→10) | Independent of pH | Hu et al. (2017) |
Polyethylene (virgin and beached) | Metals (Ag, Cd, Co, Ni, Pb, Zn) | Increase (pH 4→10) | Increased | Turner & Holmes (2015) |
Metal (Cr) | Decreased | |||
Metal (Cu, Hg) | Unclear | |||
Polyethylene, Polystyrene | Surfactant (Perfluorooctanesulfonic acid) | Decreasing | Increased | Wang et al. (2015) |
High density polyethylene | Metals (Cd, Co, Ni, Pb) | Increasing | Increased | Holmes et al. (2014) |
Metals (Cr) | Decreased | |||
Ionic Strength | ||||
Polystyrene | Polycyclic hydrocarbons (Naphthalene) | Up to 0.5 mM | Increased | Hu et al. (2020) |
5 mM – 50 mM | Decreased | |||
Polystyrene | Metalloid (Arsenic) | Increased | Decreased | Dong et al. (2020a) |
Polyethylene | Pesticides (Carbendazim, Dipterex, Malathion, Diflubenzuron, Difenoconazole) | Increased | Increased | Wang et al. (2020a) |
Polytetrafluoroethylene | Metalloid (Arsenic) | Increase (0 M→1 M) | Decreased | Dong et al. (2019) |
Polystyrene | Antibiotic (Oxytetracycline) | Increased | Decreased | Zhang et al. (2018) |
Polystyrene, Polyvinylchloride, Polypropylene, Polyethylene | Antibiotic (Tylosin) | Increase (0 M→0.1 M) | Increased | Guo et al. (2018) |
Increase (>0.1 M) | Decreased | |||
Salinity | ||||
Polyethylene, Polystyrene, Polypropylene, Polyamide, Polyvinylchloride | Antibiotics (Ciprofloxacin, Amoxicillin) | Presence of salt | Decreased | Liu et al. (2019), Li et al. (2018a) |
Polyethylene | Antibiotic (Sulfamethoxazole) | Increase (0.05%→3.5%) | Independent of salinity | Xu et al. (2018c) |
Polypropylene | Brominated flame retardants (Tris-(2,3-dibromopropyl) isocyanurate (TBC) and Hexabromocyclododecanes (HBCDs)) | Increase (0.05%→14%) | Increased | Liu et al. (2018b) |
Increase (14%→21%) | Decreased | |||
Polyethylene, Polystyrene | Lubrication oil | Increase (0.001→0.1 mol/L) | Increased | Hu et al. (2017) |
Polyethylene, Polyvinylchloride | Pesticide (DDT) | Presence of salt | Decreased | Bakir et al. (2014) |
High density polyethylene | Metals (Cd, Co, Ni) | Increasing | Decreased | Holmes et al. (2014) |
Metals (Cr) | Increased | |||
Metals (Cu, Pb) | independent of salinity | |||
Temperature | ||||
Polystyrene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Increase (15 °C→45 °C) | No effect | Li et al. (2019) |
Polytetrafluoroethylene | Metalloid (Arsenic) | Increase (15 °C→35 °C) | Decreased | Dong et al. (2019) |
Polypropylene | Brominated flame retardants (Tris-(2,3-dibromopropyl) isocyanurate (TBC) and Hexabromocyclododecanes (HBCDs)) | Increase (5 °C→15 °C) | Increased | Liu et al. (2018b) |
Increase (15 °C→45 °C) | Decreased | |||
Dissolved Organic Matter (DOM) | ||||
Polyethylene, Polypropylene, Polystyrene | Antibiotic (Tetracycline) | Addition of DOM (fulvic acid) | Decreased | Xu et al. (2018b) |
Polyethylene | Antibiotic (Sulfamethoxazole) | Addition of DOM (0→20 mg/L) | No significant effect | Xu et al. (2018c) |
Polyethylene | Pharmaceuticals and personal care products (Carbamazepine, triclosan, 17ᾳ-ethinyl estradiol) | Addition of DOM | Decreased | Wu et al. (2016) |
Type of microplastic . | Type of contaminant sorbed . | Change in parameter . | Impact on contaminant sorption on MPs . | References . |
---|---|---|---|---|
pH | ||||
Polystyrene | Metalloid (Arsenic) | Increase (pH 3→7) | Decreased | Dong et al. (2020a) |
Polyethylene | Pesticides (Carbendazim, Dipterex) | Increase (pH 2→6) | Decreased | Wang et al. (2020a) |
Pesticides (Malathion) | Increase (pH 2→3) | Increased | ||
Pesticides (Diflubenzuron, Difenoconazole) | Increase (pH 2→6) | Increased | ||
Polytetrafluoroethylene | Metalloid (Arsenic) | Increase (pH 3→7) | Decreased | Dong et al. (2019) |
Polystyrene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Increase (pH 3→6) | Increased | Li et al. (2019) |
Increase (pH 6→12) | Decreased | |||
Polystyrene, Polyvinylchloride | Antibiotic (Tylosin) | Increase (pH 3→7) | Decreased | Guo et al. (2018) |
Polyethylene, Polypropylene, Polystyrene | Antibiotic (Tetracycline) | Increase (pH 2→6) | Increased | Xu et al. (2018b) |
Polyethylene | Antibiotic (Sulfamethoxazole) | Increase (pH 2→12) | Increased (slightly) | Xu et al. (2018c) |
Polyethylene, Polystyrene | Lubrication oil | Increase (pH 1→10) | Independent of pH | Hu et al. (2017) |
Polyethylene (virgin and beached) | Metals (Ag, Cd, Co, Ni, Pb, Zn) | Increase (pH 4→10) | Increased | Turner & Holmes (2015) |
Metal (Cr) | Decreased | |||
Metal (Cu, Hg) | Unclear | |||
Polyethylene, Polystyrene | Surfactant (Perfluorooctanesulfonic acid) | Decreasing | Increased | Wang et al. (2015) |
High density polyethylene | Metals (Cd, Co, Ni, Pb) | Increasing | Increased | Holmes et al. (2014) |
Metals (Cr) | Decreased | |||
Ionic Strength | ||||
Polystyrene | Polycyclic hydrocarbons (Naphthalene) | Up to 0.5 mM | Increased | Hu et al. (2020) |
5 mM – 50 mM | Decreased | |||
Polystyrene | Metalloid (Arsenic) | Increased | Decreased | Dong et al. (2020a) |
Polyethylene | Pesticides (Carbendazim, Dipterex, Malathion, Diflubenzuron, Difenoconazole) | Increased | Increased | Wang et al. (2020a) |
Polytetrafluoroethylene | Metalloid (Arsenic) | Increase (0 M→1 M) | Decreased | Dong et al. (2019) |
Polystyrene | Antibiotic (Oxytetracycline) | Increased | Decreased | Zhang et al. (2018) |
Polystyrene, Polyvinylchloride, Polypropylene, Polyethylene | Antibiotic (Tylosin) | Increase (0 M→0.1 M) | Increased | Guo et al. (2018) |
Increase (>0.1 M) | Decreased | |||
Salinity | ||||
Polyethylene, Polystyrene, Polypropylene, Polyamide, Polyvinylchloride | Antibiotics (Ciprofloxacin, Amoxicillin) | Presence of salt | Decreased | Liu et al. (2019), Li et al. (2018a) |
Polyethylene | Antibiotic (Sulfamethoxazole) | Increase (0.05%→3.5%) | Independent of salinity | Xu et al. (2018c) |
Polypropylene | Brominated flame retardants (Tris-(2,3-dibromopropyl) isocyanurate (TBC) and Hexabromocyclododecanes (HBCDs)) | Increase (0.05%→14%) | Increased | Liu et al. (2018b) |
Increase (14%→21%) | Decreased | |||
Polyethylene, Polystyrene | Lubrication oil | Increase (0.001→0.1 mol/L) | Increased | Hu et al. (2017) |
Polyethylene, Polyvinylchloride | Pesticide (DDT) | Presence of salt | Decreased | Bakir et al. (2014) |
High density polyethylene | Metals (Cd, Co, Ni) | Increasing | Decreased | Holmes et al. (2014) |
Metals (Cr) | Increased | |||
Metals (Cu, Pb) | independent of salinity | |||
Temperature | ||||
Polystyrene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Increase (15 °C→45 °C) | No effect | Li et al. (2019) |
Polytetrafluoroethylene | Metalloid (Arsenic) | Increase (15 °C→35 °C) | Decreased | Dong et al. (2019) |
Polypropylene | Brominated flame retardants (Tris-(2,3-dibromopropyl) isocyanurate (TBC) and Hexabromocyclododecanes (HBCDs)) | Increase (5 °C→15 °C) | Increased | Liu et al. (2018b) |
Increase (15 °C→45 °C) | Decreased | |||
Dissolved Organic Matter (DOM) | ||||
Polyethylene, Polypropylene, Polystyrene | Antibiotic (Tetracycline) | Addition of DOM (fulvic acid) | Decreased | Xu et al. (2018b) |
Polyethylene | Antibiotic (Sulfamethoxazole) | Addition of DOM (0→20 mg/L) | No significant effect | Xu et al. (2018c) |
Polyethylene | Pharmaceuticals and personal care products (Carbamazepine, triclosan, 17ᾳ-ethinyl estradiol) | Addition of DOM | Decreased | Wu et al. (2016) |
Wu et al. studied the transport of some personal care products and pharmaceuticals through polyethylene. It was found that the presence of dissolved organic matter significantly reduces the sorption capacity of microplastics for these contaminants (Wu et al. 2016) (Table 2). It happens because of the competition between organic matter and contaminants to get the sorption sites on microplastics (Cox et al. 2007). Moreover, there is also a possibility of the sorption of contaminants over the organic matter (Ilani et al. 2005; Wu et al. 2016). In such a scenario, adsorption of contaminant laden organic material over the microplastics may pose further risk. Nevertheless, some environmental factors help in reducing the sorption of organic compounds over microplastics. For example, interaction of microplastic particles with oxygen enhances the surface polarity in particles, which reduces the adsorption of organic moieties (Huffer et al. 2018). Similarly, weathering processes result in increased crystallinity of microplastic particles thereby helping in reducing the adsorption (Hartmann et al. 2017).
Environmental factors not only play a significant role in the transfer of chemical contaminants through microplastics; rather, colonization of microbial communities over the surface of microplastics is also affected considerably by the changes in environmental parameters (McCormick et al. 2014; Oberbeckmann et al. 2018). One of the reasons is that plastic surfaces are comparatively more stable than other natural materials and hence provide a stable base to microbes. Further, biogeography and seasonal variation in temperature and/or salinity have an important role in deciding the type of communities colonizing over the microplastics (Fuhrman et al. 2008; Oberbeckmann et al. 2012, 2014; Amaral-Zettler et al. 2015). A summarized list of various environmental parameters affecting the sorption of contaminants over microplastics is shown in Table 2.
Effect of physico-chemical properties of microplastics on their contaminant transport capacity in drinking water matrices
Physico-chemical properties of microplastics such as chemical composition, crystal structure, size, colour, density, etc. have a profound impact on their contaminant transport ability (Table 3). In general, microplastics possess higher sorption capacity for hydrophobic contaminants (Yu et al. 2019). Hydrophobic contaminants are usually lipophilic, and hence tend to sorb more on the microplastics, compared to the hydrophilic contaminants (Takada 2006). Specifically, these are found to dominate for sorption onto the polyethylene, polystyrene, and polyvinylchloride (Wang et al. 2015). Polyaromatic hydrocarbons (PAHs) are the typical hydrophobic pollutants in the environment. Strong interaction between PAHs and micro- or nanoscale plastics has been evidenced in aquatic environment (Tan et al. 2019). Hydrophobic interaction also plays an important role in the sorption behaviour of polychlorinated biphenyls (PCBs) such as diaclor, diconal, educarel, etc. over the microplastics. It has been shown that the sorption of PCBs on polyethylene plastic films was significantly higher than that of polystyrene and polyvinyl chloride. Later, these chemical-laden microplastics may get transferred to aquatic biota and further into the food chain through ingestion, sorption, and/or respiration (Hartmann et al. 2017). Therefore, it is important to understand the interaction behaviour of microplastics as well as contaminants. The sorption isotherms of various microplastics provide insights into their sorption behaviour. Sorption isotherms of polyethylene were found to be highly linear, which indicated that the sorption behaviour tended towards sorption into the bulk polymer. However, the sorption isotherms of polystyrene were nonlinear and п – п interaction played a crucial role in the adsorption (Huffer & Hofmann 2016). The sorption isotherms of perfluorooctanesulfonic acid and perfluorooctanesulfonamide on the three different types of microplastics, viz. polyethylene, polystyrene, and polyvinylchloride, were highly linear, which revealed that the dominant interaction process was partitioning rather than hydrophobicity (Wang et al. 2015).
Type of microplastic . | Type of contaminants sorbed . | Change in parameter . | Impact on contaminant sorption on MPs . | References . |
---|---|---|---|---|
Particle size | ||||
Polystyrene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Decrease | Increased | Li et al. (2019), Wang et al. (2018) |
Polypropylene | Brominated flame retardants (Tris-(2,3-dibromopropyl) isocyanurate (TBC) and Hexabromocyclododecanes (HBCDs)) | Decrease | Increased | Liu et al. (2018b) |
Polypropylene | Fungicide (Difenoconazole) | Decrease | Increased | Goedecke et al. (2017) |
Ageing | ||||
Polypropylene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Increase | Increased | Wu et al. (2020) |
Polypropylene | Benzene, toluene, ethylbenzene, and xylene (BTEX) | Increase | Independent of ageing | Muller et al. (2018) |
Polystyrene | BTEX | Increase | Decreased | |
Polypropylene | Fungicide (Difenoconazole) | Increase | Increased | Goedecke et al. (2017) |
Resin pellets (along beaches) | Pesticide (DDT) | Increase | Increased | Antunes et al. (2013) |
Polypropylene | Polyaromatic hydrocarbon (Phenanthrene) | Increase | Increased | Karapanagioti & Klontza (2008) |
Colour | ||||
Microplastics | Polychlorinated biphenyls, Polyaromatic hydrocarbons | Black or any dark colour | Increased | Frias et al. (2013), Antunes et al. (2013) |
Density | ||||
Microplastics | Chemicals | Low density | Increased | Lee et al. (2018), Fries & Zarfl (2012), Karapanagioti & Klontza (2008) |
High density | Decreased |
Type of microplastic . | Type of contaminants sorbed . | Change in parameter . | Impact on contaminant sorption on MPs . | References . |
---|---|---|---|---|
Particle size | ||||
Polystyrene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Decrease | Increased | Li et al. (2019), Wang et al. (2018) |
Polypropylene | Brominated flame retardants (Tris-(2,3-dibromopropyl) isocyanurate (TBC) and Hexabromocyclododecanes (HBCDs)) | Decrease | Increased | Liu et al. (2018b) |
Polypropylene | Fungicide (Difenoconazole) | Decrease | Increased | Goedecke et al. (2017) |
Ageing | ||||
Polypropylene | Broad spectrum antimicrobial and endocrine disrupting chemical (Triclosan) | Increase | Increased | Wu et al. (2020) |
Polypropylene | Benzene, toluene, ethylbenzene, and xylene (BTEX) | Increase | Independent of ageing | Muller et al. (2018) |
Polystyrene | BTEX | Increase | Decreased | |
Polypropylene | Fungicide (Difenoconazole) | Increase | Increased | Goedecke et al. (2017) |
Resin pellets (along beaches) | Pesticide (DDT) | Increase | Increased | Antunes et al. (2013) |
Polypropylene | Polyaromatic hydrocarbon (Phenanthrene) | Increase | Increased | Karapanagioti & Klontza (2008) |
Colour | ||||
Microplastics | Polychlorinated biphenyls, Polyaromatic hydrocarbons | Black or any dark colour | Increased | Frias et al. (2013), Antunes et al. (2013) |
Density | ||||
Microplastics | Chemicals | Low density | Increased | Lee et al. (2018), Fries & Zarfl (2012), Karapanagioti & Klontza (2008) |
High density | Decreased |
The porous and non-porous nature of microplastics is also important to determine its interaction with the contaminants. Some microplastics are non-porous (such as polyethylene) while others may be meso-porous (such as polystyrene). The non-porous nature of polyethylene results in higher sorption capability for hydrophobic organic compounds (Wang & Wang 2018; Zuo et al. 2019). On the other hand, the high sorption affinity of polystyrene for PAHs is attributed to the amorphous structure (Rochman et al. 2013). Similarly, it has also been reported that crystalline structure of microplastics reduces the adsorption of contaminants. The reason crystalline structures resist adsorption while amorphous structures favour it lies in the fact that it takes huge amounts of energy to destabilize the well-oriented polymer chains of crystalline structures (Liu et al. 2019). Similarly, the surface polarity of microplastics is another property that has an influence. The strength of interaction of the contaminants with the microplastics can be determined by the surface polarity of the microplastics, such as hydrophobic organic contaminants adhere more to the non-polar surfaces and vice-versa (Mato et al. 2002; Fred-Ahmadu et al. 2020). Further, the high surface area and small size can also enhance the adsorption capacity owing to more availability of active sites for adsorption (Yu et al. 2019).
The density of the microplastics influences sorption behaviour as well (Wang et al. 2018). The sorption process can be evaluated in terms of diffusion coefficients. Higher diffusion coefficient results in more sorption and vice-versa. It has been demonstrated in a study that pollutants have low diffusion coefficients in high-density polyethylene while high diffusion coefficient in low-density polyethylene, resulting in higher sorption of contaminants over low-density polyethylene compared to high-density polyethylene (Mato et al. 2001, 2002; Teuten et al. 2007; Karapanagioti & Klontza 2008; Fries & Zarfl 2012; Lee et al. 2018). This is so because high-density polyethylene has minimal branching compared to low-density polyethylene making it more rigid and less permeable leading to low diffusion of contaminants (Saleem et al. 1989). It is also reported that dark-coloured microplastic particles tend to sorb more contaminants compared to their uncoloured counterparts (Antunes et al. 2013). This is because of the fact that dark coloured microplastics contain higher amount of additives, such as polyurethane, which promote sorption (Wang et al. 2018). The cellular membrane-like structure of these additives helps the contaminating molecules to penetrate deeper inside the polymeric substance, thus making it possible for them to be carried away along with the microplastics (Dmitrienko & Zolotov 2002).
Ageing is another important parameter that affects the contaminant transport ability of the microplastics (Table 3). Ageing of the microplastics results in various oxidation processes that alter the composition as well as the structure of the particles in terms of crystal structure, size, exposed surface area, etc. (Wu et al. 2020). Sorption isotherms of various contaminants over microplastic particles have shown that crystal structure gets disturbed upon ageing, resulting in a more non-crystalline domain, which requires very low energy to disturb the polymer chain to adsorb the contaminants (Liu et al. 2019). The hydrogen bond among oxygen-containing functional groups is an important interaction, which could be generated on the surface of microplastics during the ageing process; hence, the sorption capacity of aged microplastics is found to be significantly higher than that of the virgin material (Yu et al. 2019). It has also been shown that aged fragments accumulate metals to a greater extent without reaching equilibrium, while virgin plastics attain equilibrium at a faster rate, thus accumulating less (Brennecke et al. 2016). Overall, ageing has been found to enhance the contaminant transport capacity of the microplastics (Antunes et al. 2013). In contradiction to this, Muller et al. reported that ageing does not play any role in the sorption process of polypropylene (semi-crystalline), while in the case of polystyrene (amorphous) ageing results in low sorption of BTEX (benzene, toluene, ethylbenzene, and xylene) (Muller et al. 2018). The reason for low sorption of BTEX onto the aged polystyrene is reported to be the formation of an oxidized surface layer which increases surface polarity and thus, reduces the sorption of non-polar BTEX. Therefore, different microplastics may behave differently upon ageing, and chemical composition of the contaminants as well as that of the microplastics themselves plays an important role in this.
MICROPLASTICS IN DRINKING WATER: IMPLICATIONS FOR HUMAN HEALTH
Owing to the ubiquitous presence of microplastics, their entry into human beings and the consequent impact on health are inevitable. Microplastic ingestion in human may take place through contaminated drinking water (Kosuth et al. 2018), along with many other routes, for example, eating/drinking food items (Kosuth et al. 2018; Conti et al. 2020), seafood (Smith et al. 2018), honey and sugar (Liebezeit & Liebezeit 2013), commercially available common salt (Peixoto et al. 2019), and other plastic-wrapped food items (Rist et al. 2018). After ingestion, some of these particles may get excreted from human body via urine, bile, or faeces and with other metabolic functions due to their resistance to degradation (Wright & Kelly 2017). The rate of elimination of microplastics from the body is affected by the shape, size, polymer type, and additive chemicals associated with the ingested microplastics (Lusher et al. 2017). However, these particles also have the potential for bioaccumulation, especially with cumulative/chronic exposure, in secondary organs (Smith et al. 2018) after translocation from the gut. Owing to this bioaccumulation process, microplastics may prove to be harmful because of their inherent tendency of causing tissue obstruction (Peda et al. 2016). It has been shown that polystyrene microplastics do accumulate in the kidney, lungs, and intestine of mice resulting in oxidative stress, changes in lipid and energy metabolism, and neurotoxicity (Deng et al. 2017). Moreover, circulating microplastics have been shown to induce pulmonary hypertension and vascular dysfunction in in-vivo animal studies (Prata et al. 2020).
Research has also shown that ingested microplastic particles translocate out of the intestine through adsorption of particles adherent to mucus by specialized M-cells found over Peyer's patches (intestinal lymphoid tissue) (Ensign et al. 2012) or even via paracellular transfer through the single layer of the intestinal epithelium (persorption) (Prata et al. 2020). Microplastics (>700 nm) have been reported in human blood samples, which indicates the bioavailability of plastic particles in the human bloodstream (Leslie et al. 2022). Upon reaching the circulatory/lymphatic system, the microplastics are then carried to distant organ systems (Smith et al. 2018). It is important here to note that small microplastics owing to their large surface area do have higher potential for acting as adsorbate for contaminants and also for reaching up to the distant organs; thus posing comparatively serious health effects. Moreover, adsorption of plastic particles may also take place along with the large proteins in the body, which may induce alterations in the immune system (Powell et al. 2007; Sana et al. 2020).
The effect of microplastics on human health varies according to exposure characteristics and host susceptibility (Smith et al. 2018). Microplastic toxicity has mostly been studied for inhaled particles. The transport and effect of microplastic particles following ingestion has not been extensively studied (WHO 2019). Preliminary research has demonstrated several mechanisms of microplastic effect on human health such as exaggerated inflammatory response, genotoxicity, and oxidative stress resulting in cell and tissue damage, fibrosis, and potentially carcinogenesis (Deng et al. 2017; Schirinzi et al. 2017). Organ specific toxicity has been reported in the gastro-intestinal system, liver, reproductive system, and neurological system (Chang et al. 2020; Rai et al. 2021). Effect of microplastics on distant human organ systems ranges from an increased incidence of immune or neurodegenerative diseases (Prata et al. 2020), increased risk of lung diseases (Dong et al. 2020b), impairment in renal function (Monti et al. 2015) to bone loss secondary to an unopposed increase in the activity of osteoclasts responsible for bone reabsorption (Ormsby et al. 2016). It is also known that polyvinyl chloride is a proven carcinogen and causes angiosarcoma of the liver (Bolt 2005; Gennaro et al. 2008). Studies have further shown that microplastic particles can potentially cross the placental barrier as well, which may pose serious consequences on embryo development (Grafmueller et al. 2015; Ragusa et al. 2021).
Apart from the microplastics themselves, chemical additives and contaminants sorbed on to these particles might also pose serious health hazards (Ziccardi et al. 2016; Barboza et al. 2018; Rist et al. 2018). It has been reported in marine organisms that translocation of contaminants adsorbed on the microplastics into other body tissues increases with the duration of passage through the gut of the organisms (Chua et al. 2014). Similar may be the fate of microplastics in the human body as well. Moreover, the chemicals/additives used in the manufacturing process of plastics/microplastics cause various impacts upon ingestion, such as reproductive abnormalities (Swan et al. 2005; Lang et al. 2008; Swan 2008). It has been demonstrated that chemicals, such as phthalates and bisphenol A (BPA), which are commonly added in microplastics, are found in the human body (Thompson et al. 2009). Moreover, epidemiological studies have proven the relation between phthalate levels and adverse human health effects (Swan et al. 2005). Microplastics are also known to adsorb various metals/metalloids, such as cadmium, manganese, lead, arsenic, copper, zinc, chromium, etc., on their surfaces (Brennecke et al. 2016; Gao et al. 2019; Selvam et al. 2021). Polyethylene terephthalate particles have been reported to accumulate lead, cadmium, and zinc (Abbasi et al. 2020). Likewise, arsenic, cadmium, chromium, and lead were found to be associated with high density polyethylene (Holmes et al. 2012; Jinhui et al. 2019; Mohsen et al. 2019). Deleterious health impacts associated with metals are widely recognized (Table 4) (Khan et al. 2008; Rehman et al. 2017; Jain et al. 2018; Jain et al. 2019) and altered endocrine system and abrupt hormonal responses have been reported in organisms due to the effects of microplastics laden with metals/metalloids (Rochman et al. 2014). Furthermore, ingested microplastics can also serve as vectors of harmful bacteria that are adsorbed on their surface such as Vibrio spp. (Kirstein et al. 2016). Microplastics carry and release their microbial load inside the human body and thus, can potentially lead to disruption of gut microbiome, infections, and various other adverse health effects (Prata et al. 2020). A detailed information on the impacts of microplastics on human health is provided by Wright & Kelly (2017), Barboza et al. (2018), Rist et al. (2018). Summary of various human health effects of microplastics is given in Table 4.
Type of microplastic . | Particles/Chemical(s) associated . | Effects on human body due to the microplastics’ and/or associated chemicals . | References . |
---|---|---|---|
Polyethylene terephthalate | Lead, cadmium, and zinc | Cardiovascular effects, hypertension, reproductive issues, anaemia | Abbasi et al. (2020) |
High density polyethylene | Arsenic, cadmium, chromium, and lead | Increased risk of dermal, renal, pulmonary cancer, cardiovascular effects, hypertension | Jinhui et al. (2019), Mohsen et al. (2019), Holmes et al. (2012) |
Polycarbonate plastics, epoxy resins | Bisphenol A (BPA) | Adversely affects brain development leading to loss of sex differentiation in brain structures and behaviour, Suspected endocrine disrupting chemical | Rist et al. (2018), Talsness et al. (2009) |
Polystyrene for Styrofoam packaging | Styrene | Endocrine disrupting chemical | Rist et al. (2018) |
Polyethylene and polystyrene particles | – | Genotoxicity, apoptosis, and necrosis, leading to tissue damage, fibrosis, and carcinogenesis | Wright & Kelly (2017) |
Polyethylene and polystyrene microparticles | – | Cytotoxicity due to microparticles (polyethylene 3–16 μm; polystyrene 10 μm) at cell level due to oxidative stress in cerebral and epithelial human cell lines | Schirinzi et al. (2017) |
Polyvinyl chloride | Vinyl chloride | Angiosarcoma of liver | Gennaro et al. (2008), Bolt (2005) |
Polystyrene | – | Fast movement of particles (<100 nm) through endothelium in bone marrow and uptake by phagocytizing cells | Oberdorster et al. (2006) |
Polyvinyl chloride (in medical tubing) | Di(2-ethylhexyl) phthalate (DEHP) | High levels of BPA in infants | Green et al. (2005) |
Polystyrene particles | – | Inflammatory effects due to the small size (<100 nm) | Brown et al. (2001) |
Polyethylene particles | – | Particles up to 50 μm reported in lymph nodes, liver, and spleen, causing immune activation of macrophages and production of cytokines | Urban et al. (2000), Doorn et al. (1996), Hicks et al. (1996) |
Type of microplastic . | Particles/Chemical(s) associated . | Effects on human body due to the microplastics’ and/or associated chemicals . | References . |
---|---|---|---|
Polyethylene terephthalate | Lead, cadmium, and zinc | Cardiovascular effects, hypertension, reproductive issues, anaemia | Abbasi et al. (2020) |
High density polyethylene | Arsenic, cadmium, chromium, and lead | Increased risk of dermal, renal, pulmonary cancer, cardiovascular effects, hypertension | Jinhui et al. (2019), Mohsen et al. (2019), Holmes et al. (2012) |
Polycarbonate plastics, epoxy resins | Bisphenol A (BPA) | Adversely affects brain development leading to loss of sex differentiation in brain structures and behaviour, Suspected endocrine disrupting chemical | Rist et al. (2018), Talsness et al. (2009) |
Polystyrene for Styrofoam packaging | Styrene | Endocrine disrupting chemical | Rist et al. (2018) |
Polyethylene and polystyrene particles | – | Genotoxicity, apoptosis, and necrosis, leading to tissue damage, fibrosis, and carcinogenesis | Wright & Kelly (2017) |
Polyethylene and polystyrene microparticles | – | Cytotoxicity due to microparticles (polyethylene 3–16 μm; polystyrene 10 μm) at cell level due to oxidative stress in cerebral and epithelial human cell lines | Schirinzi et al. (2017) |
Polyvinyl chloride | Vinyl chloride | Angiosarcoma of liver | Gennaro et al. (2008), Bolt (2005) |
Polystyrene | – | Fast movement of particles (<100 nm) through endothelium in bone marrow and uptake by phagocytizing cells | Oberdorster et al. (2006) |
Polyvinyl chloride (in medical tubing) | Di(2-ethylhexyl) phthalate (DEHP) | High levels of BPA in infants | Green et al. (2005) |
Polystyrene particles | – | Inflammatory effects due to the small size (<100 nm) | Brown et al. (2001) |
Polyethylene particles | – | Particles up to 50 μm reported in lymph nodes, liver, and spleen, causing immune activation of macrophages and production of cytokines | Urban et al. (2000), Doorn et al. (1996), Hicks et al. (1996) |
MICROPLASTICS: REMOVAL METHODS IN DRINKING WATER SUPPLIES
Microplastics removal in water/wastewater/sewage treatment plants is of necessary concern, as these are one of the major sources of microplastics in drinking water supply chain (Okoffo et al. 2019). Though there are a variety of treatment options for various contaminants, microplastic-targeted treatment technologies are still in the nascent stage. It has been reported that microplastic concentration can be significantly reduced by ultrafiltration and reverse osmosis (Ziajahromi et al. 2017). In the conventional water treatment technology, primary and secondary treatment processes help in the removal of microplastics to a certain extent (Ma et al. 2019; Sun et al. 2019). The removal efficiency in wastewater treatment plants is calculated based upon the particle concentration of microplastics (viz. number of microplastic particles/litre) in the influent and effluent. As reported, approximately 50–98% of microplastics could be removed during primary treatment and 0.2–14% during secondary treatment (Sun et al. 2019). A membrane bioreactor is another option that directly treats the primary effluent. Talvitie et al. compared the microplastics’ removal efficiency in terms of particle concentration before and after the application of different tertiary treatment processes and reported that the membrane bioreactor had the highest removal efficiency (99.9%), followed by rapid sand filtration (97%), and dissolved air floatation (95%) (Talvitie et al. 2017). However, even after the tertiary treatment, a significant fraction of microplastics remains in the water, having a size range of 20–100 μm and 100–190 μm (Ziajahromi et al. 2017; Sun et al. 2019). A combination of secondary and tertiary treatment processes has also been useful in the removal of microplastics. Wang et al. reported the microplastics’ removal of approximately 61% using the sedimentation and coagulation techniques combined with the granular activated carbon (GAC) filtration (Wang et al. 2020b). Biochar-based filters have also been reported, which are able to remove microplastic particles of size up to 10 μm diameter (Wang et al. 2020c). As the impacts pertaining to the specific concentration and size range of microplastics re under research and anything specific is tough to say at present, research on the other technologies for removal of microplastics is needed.
Modification in various operational parameters of the existing water treatment technologies is another viable solution for microplastics removal. Ma et al. investigated the removal of polyethylene using aluminium and iron-based salts (AlCl3·6H2O and FeCl3·6H2O, respectively) and reported that aluminium-based salts perform better compared to the iron-based salts, (Ma et al. 2019). It was also found that small particle size of plastics resulted in a higher removal rate owing to a larger surface area, which gets easily captured and removed by flocs. It was observed that the ionic strength and turbidity of water did not influence the removal efficiency (Ma et al. 2019).
As far as microplastics-targeted treatment technologies are concerned, a gravity-powered filtration system was designed to be used in the secondary effluent in wastewater treatment plants (Beljanski et al. 2016). However, this system is yet to be tested for real wastewater. Microplastics removal through dynamic membranes has also been tested at the lab-scale (Li et al. 2018b). These membranes were formed on a 90 μm mesh and operated under gravity; however, the feasibility of using these modalities at large scale is questionable, considering the high operational and construction cost involved. Moreover, a detailed review on the removal techniques of microplastics from water is provided by Padervand et al. (2020) and Singh et al. (2021).
Considering these findings, it can be said that the microplastics-targeted treatment technologies need to be explored further. Besides improvements in the wastewater treatment plant technologies, other simple and effective mechanisms also need to be developed. In this context, advancement in end-use treatment options, such as efficient filtration, may prove to be advantageous. Such systems should be made to remove microplastic particles on their own without combining with any other treatment method. Research is also needed to target the removal of specific size ranges of microplastics as small size particles (submicro- and nano-range) are known to penetrate deep into the body tissues. Moreover, minimization of the microplastics’ production is the key issue. In this, enforcement of strict regulation and safe disposal of plastic products is of priority concern (Cheung & Fok 2017).
MICROPLASTICS: REGULATORY ACTIONS
Considering the seriousness of the issue of microplastic pollution, stringent actions on the global level are highly solicited. Sometimes, bioplastics are referred to as an alternative of conventional plastics considering their biodegradable properties (Hoffman et al. 2019). However, there are enough evidences to support that bioplastics do instigate a variety of risks for humans as well as other biotic species (Shruti & Muniasamy 2019). Therefore, regulations should be for the production, use, and safe disposal of microplastic particles. Though specific targeted actions are yet to be taken, various guidelines and initiatives have come up with an intent to curb and/or regulate microplastics pollution. Most of these initiatives have focussed on the minimization of production and use of (primary) microplastics, which will ultimately reduce the burden of microplastics in the environmental matrices and consequently in drinking water sources. Based on the Millennium Development Goals, United Nations General Assembly has targeted the sound management of various chemicals and other wastes with a focus on reducing their release to various environmental compartments (air/water/soil) for minimizing their effects on health as well as environment (UN 2015). The world's major seven developed countries (G7) in 2015 discussed alternatives for plastic pollution (G7 2015) and currently are in the process of developing an action plan to tackle marine waste including plastic waste. REACH (Registration, Evaluation, Authorization, and Restriction of Chemicals) regulation for controlling the use of various chemicals has been propounded by the European Union (EU) in 2006 (REACH 2006). It addresses the issues of plastic monomers and associated additives. Recently, the European Chemical Agency (ECHA) proposed varied restrictions on the use of microplastics in products available in the EU market, so that their release into the environment may be minimized (ECHA 2019). In absolute terms, the proposal asks for reducing the release of 500,000 tonnes of microplastics over 20 years.
Various other governmental agencies and industrial sectors have also come forward and adopted new guidelines to combat pollution due to microplastics. In 2012, one of the leading consumer goods company – Unilever declared to phase out microplastics by 2015 from their personal care products, such as soap, skin scrubs, shower gel, etc. (Unilever 2012). Similarly, other brands such as Johnson and Johnson, Procter & Gamble, L'oreal, Colgate-Palmolive, etc. have also decided to ban the use of microplastics in their products (Fauna & Flora 2018). To take things forward, many nations have either proposed a new set of rules or amended the existing ones to restrict the use of microplastics. For example, the United States has promulgated the Microbead-free Waters Act in 2015, which prohibits the use of plastic particles in the manufacture of various rinse-off cosmetics products such as shampoos, soaps, toothpastes, etc. (McDevitt et al. 2017; US FDA 2017). Similarly, Canada, France, New Zealand, the United Kingdom, and many others have proposed regulations for curtailing the production and use of microplastics. A summary of various global/national initiatives is shown in Table 5.
Country . | Regulation . | Product category . | Definition of microplastics . |
---|---|---|---|
Canada (in force) | Microbeads in Toiletries Regulations (Canada Gazette, Part II: Vol. 151, No. 12, 2017) | Toiletries, meaning any personal hair, skin, teeth, or mouth care products for cleansing or hygiene, including exfoliants | Microbead: plastic microbeads that are ≤5 mm in size, any plastic particle, including different forms such as solid, hollow, amorphous and solubilized. |
France (in force) | Decree prohibiting the placing on the market of rinse-off cosmetic products for exfoliation or cleansing that contain solid plastic particles (Amendment CD1857, 2019) | Rinse-off cosmetic products for exfoliation or cleansing | Solid plastic particles, with the exception of particles of natural origin not liable to persist in, or release active chemical or biological ingredients into the environment or to affect animal food chains |
Ireland (in force) | Microbeads (Prohibition) Act, 2019 (S.I. 52 of 2019) | Outlawed the sale, manufacture, import, and export of products containing microplastics | Solid particle which is water insoluble and size ranging from 5 mm–1 nm |
New Zealand (in force) | Waste Minimization (Microbeads) Regulations, 2017 (Section 23(1)(b), 2017/291) | Wash-down cosmetic products, cleaning products | Microbead: a water-insoluble plastic particle that is less than 5 mm at its widest point |
Sweden (in force) | Draft regulation prohibiting the placing on the market of rinse-off cosmetics that contain solid plastic particles which have been added for exfoliating, cleaning, or polishing purposes (Ordinance (1998:944), 4 – 4b) | Rinse-off cosmetics products | Solid particles of plastic which are 5 mm or less in size in any dimension and which are insoluble in water |
United Kingdom (England, Wales, Scotland) | The Environmental Protection (Microbeads) Regulations, 2017/18 (2017 No. 1312) | Rinse-off personal care products | Microbead: any water-insoluble solid plastic particle of less than or equal to 5 mm in any dimension |
United States (in force) | Microbead-free Waters Act, 2015 (Public Law 114–114, 114th Congress, Dec. 28, 2015) | Rinse-off cosmetics products | Microbead: any solid plastic particle that is less than 5 mm in size and is intended to be used to exfoliate or cleanse the human body or any part thereof. |
Belgium (notified) | Draft Sector Agreement to support the replacement of microplastics in consumer products (Notification Number 2017/0465/B’ (2 October 2017) | – | – |
China (notified) | ‘Opinions on further strengthening the clean-up of plastic pollution’ which laid out the plan to ban the manufacture of daily chemical products containing plastic microbeads (NDRC [2020] No. 80) | Daily chemical products | – |
India (notified in 2017, implementation since 2020) | Classification for cosmetic raw materials and adjuncts, Part 2: List of raw materials generally not recognized as safe for use in cosmetics (BIS IS 4707−2:2017) | Cosmetic products | Non-biodegradable polymeric microbeads |
Italy (notified) | Draft technical regulation banning the marketing of non-biodegradable and non-compostable cotton buds and exfoliating rinse-off cosmetic products or detergents containing microplastics (Notification No. 2018/258/I) | Exfoliating rinse-off cosmetics products and detergents | Water insoluble solid plastic particles of 5 mm or less, referring to definition in Commission Decision EU 2017/1217 of 23 June 2017 |
Japan (notified in 2018) | Enacted a bill aimed at reducing the use of microplastics, contained in some cosmetics and other products | Cosmetic and other products | – |
South Korea (notified in 2021) | Draft regulation on Safety Standards etc. of Cosmetics | Cleansing products, dental cleansing products | Microbead: less than or equal to 5 mm in size |
Country . | Regulation . | Product category . | Definition of microplastics . |
---|---|---|---|
Canada (in force) | Microbeads in Toiletries Regulations (Canada Gazette, Part II: Vol. 151, No. 12, 2017) | Toiletries, meaning any personal hair, skin, teeth, or mouth care products for cleansing or hygiene, including exfoliants | Microbead: plastic microbeads that are ≤5 mm in size, any plastic particle, including different forms such as solid, hollow, amorphous and solubilized. |
France (in force) | Decree prohibiting the placing on the market of rinse-off cosmetic products for exfoliation or cleansing that contain solid plastic particles (Amendment CD1857, 2019) | Rinse-off cosmetic products for exfoliation or cleansing | Solid plastic particles, with the exception of particles of natural origin not liable to persist in, or release active chemical or biological ingredients into the environment or to affect animal food chains |
Ireland (in force) | Microbeads (Prohibition) Act, 2019 (S.I. 52 of 2019) | Outlawed the sale, manufacture, import, and export of products containing microplastics | Solid particle which is water insoluble and size ranging from 5 mm–1 nm |
New Zealand (in force) | Waste Minimization (Microbeads) Regulations, 2017 (Section 23(1)(b), 2017/291) | Wash-down cosmetic products, cleaning products | Microbead: a water-insoluble plastic particle that is less than 5 mm at its widest point |
Sweden (in force) | Draft regulation prohibiting the placing on the market of rinse-off cosmetics that contain solid plastic particles which have been added for exfoliating, cleaning, or polishing purposes (Ordinance (1998:944), 4 – 4b) | Rinse-off cosmetics products | Solid particles of plastic which are 5 mm or less in size in any dimension and which are insoluble in water |
United Kingdom (England, Wales, Scotland) | The Environmental Protection (Microbeads) Regulations, 2017/18 (2017 No. 1312) | Rinse-off personal care products | Microbead: any water-insoluble solid plastic particle of less than or equal to 5 mm in any dimension |
United States (in force) | Microbead-free Waters Act, 2015 (Public Law 114–114, 114th Congress, Dec. 28, 2015) | Rinse-off cosmetics products | Microbead: any solid plastic particle that is less than 5 mm in size and is intended to be used to exfoliate or cleanse the human body or any part thereof. |
Belgium (notified) | Draft Sector Agreement to support the replacement of microplastics in consumer products (Notification Number 2017/0465/B’ (2 October 2017) | – | – |
China (notified) | ‘Opinions on further strengthening the clean-up of plastic pollution’ which laid out the plan to ban the manufacture of daily chemical products containing plastic microbeads (NDRC [2020] No. 80) | Daily chemical products | – |
India (notified in 2017, implementation since 2020) | Classification for cosmetic raw materials and adjuncts, Part 2: List of raw materials generally not recognized as safe for use in cosmetics (BIS IS 4707−2:2017) | Cosmetic products | Non-biodegradable polymeric microbeads |
Italy (notified) | Draft technical regulation banning the marketing of non-biodegradable and non-compostable cotton buds and exfoliating rinse-off cosmetic products or detergents containing microplastics (Notification No. 2018/258/I) | Exfoliating rinse-off cosmetics products and detergents | Water insoluble solid plastic particles of 5 mm or less, referring to definition in Commission Decision EU 2017/1217 of 23 June 2017 |
Japan (notified in 2018) | Enacted a bill aimed at reducing the use of microplastics, contained in some cosmetics and other products | Cosmetic and other products | – |
South Korea (notified in 2021) | Draft regulation on Safety Standards etc. of Cosmetics | Cleansing products, dental cleansing products | Microbead: less than or equal to 5 mm in size |
Though concerted actions are being taken at various forums, there is still a state of confusion due to the wide variations in the type of microplastics found. Specifically, a major obstacle in setting a widely applicable regulation is the lack of a universal definition of microplastics. As of now, for the sake of simplicity, any solid polymeric, non-biodegradable content having a size <5 mm is considered as microplastic (Arthur et al. 2009; Thompson et al. 2009). However, there are various issues of contradiction, for example, the small plastic particles produced in the tyre abrasion process are not considered as microplastics as per the current norms (Verschoor 2015; Brennholt et al. 2018). Moreover, there is no consensus on what should be the lower size limit for microplastics. These limits will have an impact on the actual measurement of the amount of microplastics, which is certainly a fundamental requirement for regulation purposes.
CONCLUSION AND WAY FORWARD
It is noteworthy that research on microplastics and their impact on human health is in its initial stages but still there is sufficient evidence to substantiate the enormous risks posed by these tiny particles to the environment and human health. Apart from the inhalation and ingestion of microplastics through contaminated air and food, drinking water is one of the major routes to microplastic intake. As microplastics are laden with variety of other contaminants as well, it is necessary to keep in mind the environmental as well as physico-chemical factors affecting their contaminant transport capabilities, while developing new treatment options. Though various technologies are under way to achieve microplastics’ removal from water treatment plants and the drinking water supply chain; it is better to restrict the manufacture and use of microplastics itself considering their harmful impacts posed through the intake of contaminated water and food.
Also, to understand the complete picture of the impacts of microplastics, thorough research is inevitable. As of now, most of the known adverse human health impacts caused by the microplastics are with respect to the inhaled route of exposure, while the understanding about the impacts due to the ingested route of exposure is scanty. There are several knowledge gaps in this area as well that needs to be worked upon. For example, the bio-accumulative potential of microplastics in human body is yet to be fully explored. Toxicological details such as lethal concentration and effective concentration limits also need to be defined. The impact of biofilms, pigments, microbes, and various other chemical moieties accumulated over microplastic particles also needs to be understood. Besides the scientific approach towards the reduction of these emerging contaminants, the regulatory approach needs to be targeted as well. As far as third world countries are concerned, these are still struggling to provide microbe-free safe water to their citizens; dealing with the microplastics in water might not be their priority in resource allocation. Hence, there is a need for development of a global action framework and technical as well as financial support from developed countries and multinational agencies to sensitize and combat microplastic pollution in developing countries. Though, remarkable attempts are being made to reduce the production and use of primary microplastics, it is necessary to acknowledge the fact that a significant portion of the pollution comes from the secondary sources after breakdown through abiotic and biotic means (secondary microplastics). Therefore, the target needs to be the overall management of plastic waste. The monitoring and periodic check-ups of microplastics in the surface water bodies and groundwater should be made an important component in various water quality monitoring programmes. Moreover, qualitative research involving the public health approach should also be encouraged for understanding the people's perceptions to formulate various policies.
FUNDING
The authors are thankful to the Indian Council of Medical Research (ICMR), New Delhi for the financial support (project grant number ICMR-NIREH/BPL/IMP-PJ-44/2021–22/469: Principal Investigator – Surya Singh, ICMR – NIREH, Bhopal).
CONFLICTS OF INTERESTS
The authors have no conflicts of interest to declare that are relevant to the content of this article.
DATA AVAILABILITY STATEMENT
All relevant data are included in the paper or its Supplementary Information.