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Review
Nanoplastics: Detection and impacts in aquatic environments –Areview
Nigarsan Kokilathasan, Maria Dittrich ⁎
Biogeochemistry Group, Department of Physical and Environmental Sciences, University of Toronto Scarborough, 1065 Military Trail, Toronto, ON M1C1A4, Canada
HIGHLIGHTS GRAPHICAL ABSTRACT
•We critically reviewed 150 publications
on nanoplastics in aquatic systems.
•The study identified a lack of a unified
protocol for the sampling of nanoplastics.
•Toxicological studies using natural
nanoplastics are often missing.
•Developing new techniques for the char-
acterization of nanoplastics is recom-
mended.
ABSTRACTARTICLE INFO
Editor: Dimitra A Lambropoulou
Keywords:
Nanoplastics
Aquatic ecosystems
Biofilm
Detection
Sampling
The rise in the global production of plastics has led to severe concerns about the impacts of plastics in aquatic environ-
ments. Although plastic materials degrade over extreme long periods, they can be broken down through physical,
chemical, and/or biological processes to form microplastics (MPs), defined here as particles between 1 μmand
5 mm in size, and later to form nanoplastics (NPls), defined as particles <1μm in size. We know little about the abun-
dance and effects of NPls, even though a lot of research has been conducted on theecotoxicological impacts of MPs on
both aquatic biota. Nevertheless, there is evidence that NPls can both bypass the cell membranes of microorganisms
and bioaccumulate in the tissues and organs of higher organisms. This review analyzes 150 publications collected
by searching through the databases Web of Science, SCOPUS, and Google Scholar using keywords such as
nanoplastics*,aquatic*,detection*,toxic*,biofilm*,formation*, and extracellular polymeric substance* as singular or plural
combinations. We highlight and critically synthesize current studies on the formation and degradation of NPls, NPls'
interactions with aquatic biota and biofilm communities, and methods of detection. One reason for the missing data
and studies in this area of research is the lack of a protocol for the detection of, and suitable methods for the charac-
terization of, NPls in the field. Our primary aim is to identify gaps in knowledge throughout the review and define fu-
ture directions of research to address the impacts of NPls in aquatic environments. The development of consistent and
standardized sets of procedures wouldaddress the gaps in knowledge regardingthe formation and degradation of NPls
as well as sampling and characterizing natural NPls needed to observe the full extent of NPls on aquatic biota and bio-
film communities.
Science of the Total Environment 849 (2022) 157852
⁎Corresponding author.
E-mail address: m.dittrich@utoronto.ca (M. Dittrich).
http://dx.doi.org/10.1016/j.scitotenv.2022.157852
Received 13 May 2022; Received in revised form 13 July 2022; Accepted 1 August 2022
Available online 6 August 2022
0048-9697/© 2022 Published by Elsevier B.V.
Contents lists available at ScienceDirect
Science of the Total Environment
journal homepage: www.elsevier.com/locate/scitotenv
Contents
1. Introduction................................................................ 2
2. Methods.................................................................. 3
3. Resultsanddiscussion............................................................ 4
3.1. Fateofnanoplasticsinaquaticsystems................................................. 4
3.2. Formationanddegradationofnanoplasticsinaquaticenvironments .................................... 4
3.2.1. Formationofnanoplastics................................................... 4
3.2.2. Degradationofnanoplastics.................................................. 7
3.3. Biofilmcommunitiesonnanoplastics.................................................. 7
3.4. Impactsofnanoplasticsonaquaticbiota ................................................ 8
3.5. Currentmethodsofsamplingandcharacterizingnanoplastics...................................... 10
3.5.1. Samplingmethods..................................................... 10
3.5.2. Visualizationmethods................................................... 11
3.5.3. Analyticalmethodsforcompositionanalysis.......................................... 12
4. Futureworkandconclusion........................................................ 14
CRediTauthorshipcontributionstatement .................................................... 14
Declarationofcompetinginterest........................................................ 14
Acknowledgments............................................................... 14
AppendixA. Supplementarydata....................................................... 14
References.................................................................. 14
1. Introduction
The global production of plastics reached around 368 million metric
tons in 2019, with China accounting for roughly one-quarter of that amount
(Garside, 2020). Up to 94 % of plastics generated globally are expected to
end up in landfills or be released into the environment through processes
like surface runoff and wastewater treatment plant (WWTP) effluent
(Alimba and Faggio, 2019;Alimi et al., 2018).
In aquatic environments, plastics degrade down to smaller particles
until finally reaching micro- and nano-sized particles. Theterm microplastics
(MPs) was first coined in 2004 by Thompson et al. to describe the abundance
of microscopic plastic debris observed in marine environments (Alimba and
Faggio, 2019;Thompson et al., 2004). Nano-sized plastics, or nanoplastics
(NPls), were initially observed in the environment as a product of the
photodegradation of marine MPs sampled in 2014 under controlled and en-
vironmentally relevant conditions in the laboratory (Gigault et al., 2016). In
2015, researchers found NPls within the colloidal fraction of seawater sam-
pled from the North Atlantic Subtropical Gyre (Ter Halle et al., 2017). The
observed formation and presence of NPls in the environment by Gigault
et al. (2016) and Ter Halle et al. (2017) fueled the rise in publications relat-
ing to NPls (Fig. 1a). The use of the keyword combinations polymer
nanoparticle* and plastic nanoparticle* can be traced back to 2008. We
applied these keywords to include nanotoxicity-focused studies and ag-
gregation studies featuring man-made synthetic particles (Lundqvist
et al., 2008). Despite numerous studies on NPls, there is still no consen-
sus on the definition of NPls. The definition of NPls ranges from plastic
debris smaller than 100 nm to particles between 1 nm and 999 nm (Alimi
et al., 2018;Mattsson et al., 2015;Ter Halle et al., 2017). In this review,
we refer to NPls as plastic debris <1μm in size and MPs as plastic debris be-
tween 1 μmand5mminsize(Ter Halle et al., 2017).
MPs (i.e., microbeads, films, and powders) are ubiquitous in most
aquatic environments, with the most common form being microfibers
(Akdogan and Guven, 2019;Alimba and Faggio, 2019;Elgaranhy et al.,
2021). Microfibers constitute over 95 % of total MPs in remote areas such
as the waters of Antarctica and the Arctic and the Qinghai-Tibet Plateau
(Akdogan and Guven, 2019;Alimba and Faggio, 2019;Browne et al.,
2011;Li et al., 2018). Through the ingestion and entanglement of plastic
debris, MPs cause broad physical and ecotoxicological impacts, including
starvation and suffocation (Li et al., 2018). The physicochemical properties
of MPs and NPls –for example,a large surface area, crystallinity, andhydro-
phobic surface –can promote the sorption and subsequent release of con-
taminants and additives, i.e., plasticizers and flame retardants (Elgaranhy
et al., 2021). Thus, MPs and NPls can serve as potential vectors for
the exposure of the contaminants. As the human population grows, the
abundance of MPs and NPls released into aquatic environments will in-
crease as a by-product of human activities (i.e., the production of cosmetic
products) (Elgaranhy et al., 2021).
Research on the impacts of plastics in aquatic systems has been focused
mostly on MPs: of the articles published in 2021, a total of 3006 focused on
MPs while 477 focused on NPls (Fig. 1b) (Alimba and Faggio, 2019;Alimi
et al., 2018).Although the ecological and toxicological impacts of MPs have
been widely studied in recent years, we know little about NPls. The surface-
to-volume ratios of NPls are large: a normal plastic shopping bag broken
down into NPls with a diameter of 40 nm would theoretically have an over-
all surface area of 2600 m
2
, roughly the total surface area of approximately
14,521 basketballs (Mattsson et al., 2015). The large surface area of NPls
and the presence of oxygen-containing functional groups (i.e. carbonyl
and carboxyl) from photodegradation allow for both the sorption
anddesorptionofhydrophilic organic pollutants (Duan et al., 2021;
Fu et al., 2021).Thehydrophobicforcebetweentheplasticpolymer
and the pollutant, and the π−πinteractions, strongly influences the
sorption and desorption processes of hydrophobic organic pollutants on
MPs and NPls (Agboola and Benson, 2021;Luo et al., 2022;Trevisan
et al., 2020).
Their small size also enables NPls to bypass biological barriers, such as
cell surfaces, and accumulate in tissues and organs. Furthermore, the
desorption of contaminants would elicit additional toxicity to aquatic
biota. We expect NPls to impact microbial settlement, as well as impact
the resulting biofilms on the surface of plastic debris. Biofilms are aggre-
gates of microbes embedded within a matrix comprised of extracellular
polymeric substances (EPS) in biofilms and can become entrapped
(Rummel et al., 2017;Zettler et al., 2013). Thus, NPls can attach to EPS
in biofilms and become trapped. NPlscan impact aquatic biota and humans
in ways that MPs and macroplastics cannot.
However, the lack of a unified protocol for the detection of NPls in
aquatic environments results in the inability to compare and evaluate
their abundance on a global scale (Nguyen et al., 2019;Strungaru et al.,
2019). The inability to evaluate the abundance of NPls inthe environment
is evident in the number of publications relating to the detectionand forma-
tion of NPls. For example, for the year 2021, the number of publications
yielded from the keywords combinations “nanoplastic* AND detect*”,as
well as “nanoplastic* AND form*”, were 123 and 97, respectively (Fig. 1c).
In addition to the limitations regarding the sampling of NPls, there are
several drawbacks regarding the visualization and analysis of NPls. The
NPls' low density and conductivity complicate the visualization of NPls by
electron microscopy (Jakubowicz et al., 2021;Nguyen et al., 2019). Atomic
force microscopy (AFM) can produce images of non-conductive materials
at a nanometric scale, making it ideal for visualizing NPls (Khulbe and
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
2
Matsuura, 2000;Mariano et al., 2021). However, the visualization of NPls
can become complicated when NPls form aggregates with natural colloids
through electrostatic and hydrophobic interactions between surface func-
tional groups and EPS (Besseling et al., 2017;Chen et al., 2018;Okshevsky
et al., 2020). The interactions between NPls and natural colloids not only in-
fluences the rate of sedimentation but also makes it visually and analytically
challenging to differentiate between the carbon structures of NPls with the
structures of natural colloids (Gigault et al., 2021). This potential for aggre-
gation impedes the modeling of fate and transport processes.
Conventional Fourier-transform infrared spectroscopy (FTIR) and
Fourier-transform infrared microscopy-spectroscopy (micro-FTIR) are not
directly capable of analyzing NPls because of the limitation in spatial reso-
lution. The best resolution with conventional FTIR is 1.5 μmto10μm, a size
range too large to detect NPls (Jakubowicz et al., 2021;Kurouski et al.,
2020). A minimum sample thickness of ~150 nm and deposition onto an
IR-transparent substrate are also required to avoid the complete absorption
of the infrared beam (Fu et al., 2020;Nguyen et al., 2019). Therefore, this
technique is not directly applicable for the characterization of NPls, espe-
cially those with a diameter <150 nm. Raman spectroscopy can provide
the chemical information of NPls at a better spatial resolution than FTIR
(i.e., >100 nm to ~500 nm in size), but suffers from low sensitivity and
possible auto-fluorescence backgrounds. The fluorescence background
can overshadow the Raman peaks and complicate the identification process
(Fang et al., 2020;Ghosal et al., 2018;Kumar et al., 2019).
Moreover, Raman spectroscopy is sensitive to any sorbed additives and/
or contaminants (Fu et al., 2020). Thus, in combination with the diffraction
limit of the laser spot, the analysis of NPls would yield a lot of false-positive
and false-negative signals within the Raman spectrum (Fang et al., 2020).
Surface-enhanced Raman spectroscopy (SERS) and tip-enhanced Raman
spectroscopy (TERS)are variations of Raman spectroscopy that can greatly
enhance the Raman signals at a higher spatial resolution (Verma, 2017).
Still, SERS is diffraction limited and prone to non-uniform enhancements
of the signals (Zhou et al., 2021). Even though it is a promising technique
for analyzing NPls, the metallic TER probes used in TERS have poor yields
and short lifespans (Kurouski et al., 2020). Furthermore,only one study has
reported the application of TERS in the field of NPls (Yeo et al., 2009).
Pyrolysis gas chromatography–mass spectrometry (py-GC–MS) benefits
from the reduced need to separate NPls from the environmental matrix
and ability to measure small masses of NPls (to ~50 μg). However, py-
GC–MS is sensitive to impurities and unsuitable for plastic polymers with
polar subunits and particle size analysis (Jakubowicz et al., 2021;Nguyen
et al., 2019). Due to the lack of a set of procedures to sample and analyze
NPls in aquatic systems, there is no definite way of detecting NPls in places
such as lakes that serve as a primary source of drinking water.
NPls are an emerging threat that have an extensive distribution and that
elicit toxicological impacts on aquatic biota and humans in a way that MPs
cannot in aquatic systems. Thus, it is essential to synthesize and critically
evaluate the current literature on NPls in the form of a literature review,
in order to identify current gaps in knowledge and subsequently define
future research directions. First, we examined the formation and degrada-
tion of NPls in aquatic environments. Next, we provided a critical overview
of the impacts of NPls on aquatic biota and biofilm communities. Finally,
we examined and discussed the current methods used for the separation,
visualization, and characterization of NPls.
2. Methods
A comprehensive literature review was conducted using the Web of
Science, Google Scholar, and SCOPUS databases. The main guidelines
specified in the Preferred Reporting Items for Systematic Reviews
and Meta-Analyses (PRISMA) were used (Moher et al., 2010). Between
September 2020 and April 2022, a search was based on the keywords
nanoplastic*, aquatic*, detection*, toxic*, biofilm*, ecology*, and formation*
as singular or plural combinations. In addition, publications found during
the Web of Science, Google Scholar, and SCOPUS search by including the
keywords microplastic*, extracellular polymeric substance*, EPS*,polystyrene
nanoparticle*, and polymer nanoparticle* were included, depending on
their significance and relevance to the topic of NPls. Reference lists within
the searched literature were consulted and articleswere included based on
their relevance. In total, we collected 3087 publications using the PRISMA
method. The removal of duplicates resulted in a collection of 1981 articles.
The articles were then screened based on their title, abstract and such, and
1181 publications were removed as a result. Further restrictions were also
placed on the search results: for the consistency of the search, we excluded
unpublished results or articles that were not peer reviewed from this
a)
b)
c)
Fig. 1. The number of publications appeared inscientificjournalsinthelast15years
in the Web of Science, Google Scholar, and SCOPUS databases containing keywords
a) relating to nano-sized plastic particles; b) to “nanoplastic*”or “microplastic*”;and
c) keyword combinations relating to nano-sized plastic particles. The search
results were sourced from SCOPUS and conducted on March 24, 2022.
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
3
survey. Furthermore, paperswith a publicationdate of six years or less were
favored to provide an updated review of NPls, but papers that were more
than six years old were also selected depending on their significance. Pa-
pers that were solely about MPs were excluded unless they provided com-
plementary information on NPls. Overall, 150 papers were critically
reviewedfor this study. Some literature may have been omitted, but we be-
lieve that the selected papers cover research in journals with considerable
impact (Gao et al., 2021;Haegerbaeumer et al., 2019).
Data used for the Raman spectra were retrieved from Munno et al.
(2020) through their SLoPP and SLoPP-E spectra library (Munno et al.,
2020). In particular, data was retrieved for four different plastic polymers
sampled from the waters of San Francisco, USA: PS –polystyrene (clear
beads), PP –polypropylene (clear fragments), PET –poly(ethylene
terephthalate) (clear fragments), and PE –polyethylene (clear fragments).
3. Results and discussion
3.1. Fate of nanoplastics in aquatic systems
After entering terrestrial and aquatic ecosystems, plasti cs undergo trans-
formations (Duan et al., 2021). In the topsoil, plastic debris is depredating
due to direct exposure to ultraviolet (UV) radiation, high temperature and
its fluctuations, and the presence of microbes and terrestrial organisms
(He et al., 2018). Compared to terrestrial systems, the degradation of plastic
debris is less intensive in aquatic environments since water can serve as a
heat buffer and there are fewer instances of mechanical damage (Iñiguez
et al., 2018). Nevertheless, plastics are transformed into MP and then into
NPls. NPls, belonging to the plastics family, are assumed to be persistent
in aquatic environments. The processes that impact the presence of NPls
in the systems are degradation and aggregation of NPls to larger particles;
both pathways are not well understood. The changes in the surface proper-
ties due to interactions with solutes or solids, for example the sorption and
desorption of contaminants, particles, and minerals will impact the degra-
dation, aggregation, and transport of NPls. On the other side, the degrada-
tion, aggregation, and transport of NPls depend on the environmental
properties of the surrounding media, either liquid or solid.
The physical and chemical properties of the liquid medium (i.e., salinity
and temperature), determine the sorption of contaminants on plastics
(Alimi et al., 2018). The primary mechanisms of organic pollutant sorption
on NPls include hydrophobic interactions, partition effect, and electrostatic
interactions (Fu et al., 2021). The sorption capacity of plastics is dependent
on the type of plastic polymer, as they differ in several factors, such as po-
larity, crystallinity, and the presence of certain functional groups. The sorp-
tion capacity can be increased as a result of further degradation of NPls due
to the high surface-to-volume ratio and surface oxidation (Alimi et al.,
2018;Paul et al., 2020;Piccardo et al., 2020). The generation of oxygen-
containing groups caused by weathering and aging processes increases
the polarity, roughness, and porosity of the plastic particles, thus influenc-
ing their sorption capacity (Yu et al., 2019). Yet, the extent of the contam-
inants' sorption on NPls remains unclear in terms of their environmental
and health impacts (Alimi et al., 2018). Additionally, recent studies on
the mechanisms of contaminant sorption on NPls have been done in the
lab with pristine NPls. This is not representative of the environmental con-
ditions, where the combined effects of multiple factors can accelerate or in-
hibit the sorption rate of contaminants on NPls. Therefore, further work is
needed to examine the mechanisms of contaminant sorption on NPls and
the subsequent impacts on aquatic biota under environmental conditions.
NPls can be exposed to active degradation processes like physical degra-
dation and photodegradation (Kooi et al., 2018). In the water column, oxy-
gen and hydroxyl radicals (OH∙) on the plastic surface playa prominent role
in the depolymerization of plastic polymers into individual monomers
(Duan et al., 2021). Within sediments, the degradation processes of NPls
are strongly influenced by porewater biogeochemical conditions (i.e., pH,
ionic strength, natural organic matter content, microbial community
composition, and geological background), the properties of the plastics
(e.g., polymer composition), and the physical characteristics of the
sediments (i.e., density and particle size distribution) (Alimi et al., 2018).
For example, natural organic matter (NOM) can influence the degradation
of NPls through the production of free radicals (Zhang et al., 2022). Fulvic
acid oxidized PS-NPls of 200 and 1000 nm diameters within 1 day under
light irradiation. More details on the degradation and formation of NPls
will be discussed in the next section (Section 3.2).
Environmental factors may impact and change the shape, chemical
composition, surface charges, andreactivities of NPls.The pH of the aquatic
environment also impacts the surface functional groups and surface charge
of NPls, either supporting or hindering aggregation (Wang et al., 2021).
Based on the pH of the environment, the charged NPls can interact with
one another or undergo electrostatic repulsion (Fig. 2)(Sharma et al.,
2021). Salinity influences the stability of NPls through aggregation via elec-
trostatic interactions. The aggregation of positively and negatively charged
PS-NPls was reported for the case when NPls were exposed to a salinity
(NaCl) >5 %; the hydrodynamic diameter of NPls increased drastically as
NaCl concentrations increased (Wu et al., 2019). This irreversible aggrega-
tion was attributed to the electrostatic interactions between the surface
charges of the NPls and NaCl, as the dominant van der Waals interactions
reduce the effective repulsion between NPls and NaCl (Singh et al.,
2019). Additionally, the compression of the electric double layer and
charge shielding on the surface of NPls can reduce the repulsive interac-
tions among NPl particles (Wang et al., 2021). Saline conditions can also in-
fluence the degradation processes experienced by NPls. Interestingly, the
changes in salinity can lead to the formation of salt crystals on the NPls
surfaces, and decrease the adsorption efficiency of light and, thus, the
photodegradation rate (Duan et al., 2021).
Since natural colloids are often negatively charged, cluster building
with positively charged NPls is favorable (Li et al., 2020). The aggregation
of negatively charged NPls can be induced under high concentrations of
electrolytes like Ca
2+
through cation bridging effects (Li et al., 2020). In
contrast, the presence of humic and fulvic acids, a major component of
dissolvedorganic matter (DOM),reduced the stability of NPl aggregates de-
pending on salinity (Singh et al., 2019). Hydrophobic interactions between
DOM and NPls can result in steric and electrostatic repulsion among other
NPls, thereby reducing the rate of aggregation and increasing their poten-
tial distribution range. With clay colloids, nanoplastic heteroaggregates
may form multilayer adsorption (Singh et al., 2019). Microbial colonization
on plastic surfaces is ubiquitous in aquatic environments. The negatively
charged algae Chlorella sp. was foundto form aggregates with PS-NPls mod-
ified with amino groups due to charge neutralization (Thiagarajan et al.,
2019). Cellulose from algal cells can also initiate aggregation with NPls in
solution (Bhattacharya et al., 2010;Wang et al., 2021).
The transport of NPls within the aquatic system depends on the
aggregation state of the nanoparticles and the properties of the media.
Once NPls end up in aquatic ecosystems, wind and ocean currents distribute
them away from their point of origin (Strungaru et al., 2019). NPls depos-
ited or formed within sediments, part of aquatic ecosystems, may be resus-
pended when the flow velocity is high enough to exceed the critical sheer
stress, or NPls may remain buried in sediments. Aggregates of plastics and
other particles with natural colloids are widely distributed in aquatic sys-
tems (Gigault et al., 2021). Aggregation of NPls with natural colloids,
i.e., NOM and clays, can lead to suspension and faster sedimentation in
the water column (Besseling et al., 2017). The incorporation of NPls to ma-
rine snow —organic-rich aggregates containing particulate organic matter
(POM) and inorganics —would theoretically increase the sedimentation
rate of the plastic particles (Porter et al., 2018), yet no significant difference
in sinking rates was observed between marine snow containing solely NPls
and MPs (Summers et al., 2018). The sinking rates were dependent solely on
the size of the sinking aggregates and not on the plastic containedwithin.
3.2. Formation and degradation of nanoplastics in aquatic environments
3.2.1. Formation of nanoplastics
While 95–99 % of released MPs and NPls in the U.S. are retained within
sewage sludge and removed by surface skimming or microfiltration in
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
4
WWTPs, an estimated eight trillion pieces of primary MPs still enter aquatic
environments per day (Alimi et al., 2018;Carr et al., 2016;Rochman et al.,
2015). Microfibers, released after the washing of clothes and the fabrication
of textiles, comprise a major component of the 1–5 % of MPs and NPls that
are released into aquatic environments due to the inability of WWTPs to
effectively retain microfibers (Besseling et al., 2017;Browne et al., 2011).
Other sources of the 1–5 % of primary MPs and NPls include industrial
abrasives and accidental spills, where surface runoff plays a dominant
role in the transport of MPs to aquatic systems. Primary NPls can also be re-
leased after the use of personal care products(Alimi e t al., 2018;Hernandez
et al., 2017). Secondary MPs and NPls originate from the in-situ fragmenta-
tion and degradation of plastics by physical weathering, photodegradation,
thermal degradation, and oxidation (Fig. 3)(Alimba and Faggio, 2019;
Amaral-Zettler et al., 2020). Physical weathering by abrasion against
rocks or other rough surfaces by wave action can cause the fragmentation
of MPs into NPls (Yu et al., 2019). The conversion of oxygen-containing
functional groups along plastic surfaces and in NOM to free radicals
(i.e., hydroxyl and peroxyl radicals) can occur under UV radiation, leading
to the depolymerization of plastic polymers and the subsequent formation
of NPls (Alimi et al., 2018;Yousif and Haddad, 2013;Zhang et al., 2022).
NPls can also occur via thermal degradation and oxidation as high tem-
peratures slowly degrade polymers into individual monomers (“NPls”)
(Amaral-Zettler et al., 2020).
Weathering processes can promote the formation of oxygen-containing
functional groups on the plastic surface (Duan et al., 2021). The presence of
oxygen-containing functional groups increases the susceptibility of plastics
to photodegradation. During the photodegradation process, plastic poly-
mers are depolymerized into individual monomers through the scission of
macromolecular chain bonds by free radicals (Alimi et al., 2018). The for-
mation of radicals such as polystyryl and its subsequent conversion into
peroxyl radicals was reported when PS was exposed to oxygen (Yousif
and Haddad, 2013). The photodegradation of PE was also reported as a re-
sult of the reaction between carbon-based alkyl radicals and oxygen to form
peroxyl radicals (Chamas et al., 2020). The peroxyl radicals can then un-
dergo further degradation to form carbonyl functional groups. The polar
functional groups are more susceptible to physical degradation
processes and thereby create NPls, as evidenced by the observation of
surface cracks on low-density polyethylene (LDPE) bars and the subse-
quent formation of LDPE-NPls (Amaral-Zettler et al., 2020;Menzel
et al., 2022). Both mechanical degradation and photodegradation
led to the formation of NPls from the degradation of PS disposable
coffee cups (Lambert and Wagner, 2016). Additionally, sorbed additives
can influence the photo-aging process of plastics. Phosphite antioxidant
Irgafos 168 was added to pure PP-MPs and then exposed to UV irradia-
tion in artificial seawater (Wu et al., 2021). Photodegradation, which
decreased as the concentration by weight (wt%) of Irgafos 168 in-
creased, was attributed to the prevention of hydroxyl radical formation.
Other formation pathways of NPls include thermal oxidation and
degradation. Thermal oxidation leads to a slow oxidative breakdown of
the plastic polymers (Amaral-Zettler et al., 2020). The process often occurs
at moderate temperatures and in the presence of infrared radiation. Ther-
mal degradation of nylon and (PET) has been suggested to cause the forma-
tion of NPls from plastic tea bags steeped at temperatureshigher than 95
°
C
(Chamas et al., 2020;Hernandez et al., 2019). NPls may also form through
polymer hydrolysis, especially under high water temperatures and in the
absence of oxygen (Amaral-Zettler et al., 2020). Sub-100 nm NPls were
formed after the exposure of 100
°
C ultrapure water to low-density PE
(LDPE) lined hot beverage cups and nylon food-grade bags (Zangmeister
et al., 2022).
Biological degradation is the process of decomposing organic sub-
stances within plastics through the presence and/or activities of microbes.
There are a few known microbial species that can break down plastics
(Yuan et al., 2020). The biodegradation of macroplastics and MPs can be
initiated through processes such as biofouling, the degradation of plas-
ticizers, and induced hydrolysis through the secretion of enzymes and
secondary metabolites (Lehner et al., 2019;Rummel et al., 2017). The
exposure of MPs to biofilms can change the physical properties of
MPs, such as crystallinity and stiffness; the physical changes can then
accelerate biodegradation (McGivney et al., 2020). Additionally, en-
zymes and secondary metabolites secreted from biofilm can degrade
Fig. 2. The stability and movement of NPls within the watercolumn can be attributed to different factors. Factors such as NPls, microbes, natural colloids, clays, anions, and
cations can promote or hinder aggregation through electrostatic interactions or repulsion, cluster building, van der Waals interactions, and hydrophobic interactions.
Adapted from Duan et al. (2021).
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
5
plastics (Amaral-Zettler et al., 2020). For example, PETase, MHETase,
and ISF6-4831 protein found in the bacterium Ideonella sakaiensis,de-
grade PET to monomers (Fig. 4)(Austin et al., 2018;Yoshida et al.,
2016;Yuan et al., 2020). Cyanobacteria Anabaena spiroides was shown
to degrade low-density PE (LDPE) MPsasaresultoftheirmetabolic
activities (Amobonye et al., 2021). The microbial and algal communities
within biofilms can influence the rate of degradation. Diatom colonization
of MPs was suggested to promote the colonization of hydrocarbon-
degrading bacteria onto the plastic particles (Dudek et al., 2020).
Further work is required to determine the lifespan rate of NPls in fresh-
water and marine systems. The formation rate and resulting size of NPls are
dependent on the size, polymer type, and polarity of the parent MPs and
macroplastics. After reacting with free radicals (i.e., hydroxyl radicals),
the oxygen-containing functional groups at the surfaces of larger MPs are
converted into polar functional groups like carbonyl groups, which are
more susceptible to degradation processes such as biodegradation from mi-
crobial colonization, further photodegradation, and physical degradation
(Chamas et al., 2020;Rummel et al., 2017). A larger surface area would
Fig. 4. The biodegradation of PET by PETase. PETase catalyzes the depolymerization reaction of PET to 2-hydroxyethyl terephthalic acid (MHET), bis(2-hydroxyethyl)
terephthalate (BHET), and terephthalic acid (TPA).
Modified based on Amaral-Zettler et al. (2020),Austin et al. (2018) and Yoshida et al. (2016).
Thermal
Oxidation
Direct
Photolysis Photodegradation
Polymers
Oligomers
n
Depolymerization
Monomers
Physical
Degradation
Fig. 3. Floating plastic debris are subjected to different types of degradation (depolymerization) processes to form NPls (monomers). Plastics can undergo physical
degradation by wind or wave action and by abrasion against rocks or other rough surfaces. Under UV radiation, the photodegradation of plastic polymers can be initiated
by the conversion of oxygen-containing functional groups into free radicals (i.e., hydroxyl and peroxyl radicals). The radicals can then be degradedtoformcarbonylfunc-
tional groups, thereby weakening the structureof the plastics and promoting the formation of NPls. Thermal degradation and oxidationoccur under high water temperatures
and the presence of infrared radiation.
Adapted from Amaral-Zettler et al. (2020).
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
6
lead to a higher abundance of oxygen-containing functional groups and,
thus, radical formation and the rate of photodegradation would be acceler-
ated. However, it is still unclear whether different plastic polymers (in re-
spect to chemical structure) of similar sizes and surface areas behave
differently. Depending on the surface properties, the formation of NPls
by physical, chemical, and/or biological degradation can be accelerated
or inhibited.
3.2.2. Degradation of nanoplastics
There is limited research on the degradation of NPls, even though the
same degradation processes that produce MPs lead to further degradation
of NPls to produce organic compounds. For example, indirect photolysis
in the presence of reactive transient species, e.g., hydroxyl radicals, causes
the degradation of PS-NPls until carboxylic acids and aromatic compounds
(Bianco et al., 2020). Similar degradation products were observed in direct
photolysis when PS-NPls were exposed to sunlight. In this case, the forma-
tion of formic acid and dissolved organic compounds (DOC), a component
of dissolved organic matter (DOM), were observed. The degradation of
PET-NPls can be caused by the application of the thermophilic enzyme
cutinase TfCut2 (Vogel e t al., 2021). Through the use of isothermal titration
calorimetry (ITC) in combination with thermo-kinetic models, the enzy-
matic cleavage of the ester bonds in PET-NPls was determined (Vogel
et al., 2021).
As with MPs, the photodegradation of NPls can occur through the ad-
sorption of UV radiation. This results in the formation of complexes with
conjugated unsaturated hydrocarbons (Lehner et al., 2019;Müller et al.,
2018).
14
C-PS-NPls in nanopure water were exposed to UV radiation
(254 nm), forming cracks on the NPls after 48 h (Tian et al., 2019). The cre-
ation of the cracks may be the result of the oxidation of active functional
groups weakening thestructural integrity of the NPls. The smaller particles
would then be further degraded by the presence of free radicals on the NPl
matrix. Several gaps knowledge exist regarding the degradation of NPls,
due to the potential aggregation with natural colloids and the limitation
in spatial resolution. The degradation of NPls would yield bioavailable
carbon-based organic compounds, contributing to the cycling of organic
matter. In addition, the likelihood of the sorption of organic contaminants
released by the degradation of NPls would increase, leading to additional
toxicological impacts. Thus, further investigation is warranted.
3.3. Biofilm communities on nanoplastics
Microorganisms can colonize on plastics (i.e., PS, high-density PE
(HDPE), PP, and polymethylmethacrylate (PMMA)) through attractive
and repulsive interactions, i.e., surface hydrophobicity, electrostatic
interactions, polymer topography, and molecular flexibility, between
their surfaces and the medium (Fig. 5)(Nava and Leoni, 2021;Rummel
et al., 2017). After colonization, microbes often form biofilms, defined
as aggregates of microbes embedded within a matrix composed of
EPS. Besides being responsible for keeping microbial cells in biofilm to-
gether, EPS adhere to substrates (Chen et al., 2019;Dudek et al., 2020).
Thus, EPS can interact with MPs and NPls, leading to the attachment of
microorganisms. MPs and NPls may homo- or hetero-agglomerate with
different particles and microorganisms accordingly to their size and
hydrophobicity (Gigaultetal.,2021). The aggregation can entrap
the plastic particles and facilitate the trophic transfer of MPs and NPls,
especially by filter-feeding bivalves (Rai et al., 2021;Ward and Kach,
2009).
Settlement on plastics can yield increased access to nutrients, but at a
cost of more susceptibility to predators (Amaral-Zettler et al., 2020). The
components and structural integrity of the biofilm matrix accelerate the
rate of resource capture,enzyme retention time, and tolerance of and/or re-
sistance to antibiotics or pathogens (Flemming et al., 2016). The buoyancy
of the plastic particles may also change due to density change from the
biofilm development (“biofouling”)(Chen et al., 2019). The settling rate
is dependent on the particle size and density, with larger particles settling
faster than smaller particles, and also affects the vertical fluxes of plastics
(Kooi et al., 2017). Biofouling can enable the transport of non-native spe-
cies and influence nutrient cycling, leading to eutrophication, the contami-
nation of drinking water, and the destruction of food webs. Furthermore,
many harmful algal bloom (HAB)-forming cyanobacteria also slow down
the sedimentation of biofouling plastics through the use of gas vacuoles to
increase their buoyancy, thereby increasing the duration and range of im-
pact (Yokota et al., 2017).
Plastic surfaces may be modified by the environment or organisms
through biofouling and the adsorption and desorption of organic and inor-
ganic compounds, resulting in the formation of an environmental corona or
“eco-corona”(Lespes et al., 2020). The presence of the eco-corona changes
the physical and chemical properties of the plastics, such as masking surface
properties, secreting plastic-degrading or plastic-modifying enzymes, and
degrading additives (Wang et al., 2020c;Yuan et al., 2020). The organic
particles released by the biodegrad ation of plastics create favorable con-
ditions for photo- and heterotrophic organisms to persist (Eich et al.,
2015). This addition of organic particles increases the amount of DOC
in the environment that is converted from DOM to POM via photo-
oxidation (Romera-Castillo et al., 2018). The subsequent leaching of
DOC from plastics can then induce microbial activity and the cycling
Fig. 5. Microbial colonization can occur on the surface of plastics like MPs through attractive and repulsive interactions, leading to the formation of biofilm. An eco-corona
can form on plastic surfaces as a result of biofilm development and the adsorption and desorption of organic and inorganic compounds.
Adapted from Junaid and Wang (2021) and Rummel et al. (2017).
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
7
of carbon in aquatic ecosystems by creating “hot spots”of high DOC
concentrations. More research is needed on the influence of the biofilm
matrix and EPS molecules on eco-corona formation on plastic debris.
Biofilms on plastic debris contain complex microbial communities rep-
resenting several trophic levels (i.e., primary producers and primary and
secondary consumers) and interactions (Kettner et al., 2019;Yokota et al.,
2017). Notably, the physicochemical properties of the plastic surfaces,
such as hydrophobicity, surface roughness, and electrostatic interactions
between the functional groups on the plastic surface and the cell wall,
can determine the presence and dominance of certain bacterial species
within the biofilm (Lagarde et al., 2016;Rummel et al., 2017). The bio-
film community on PVC, was more distinct than the biofilm community
on HDPE, LDPE, and PET under dim light conditions after 1 week of in-
cubation (Pinto et al., 2019).
In brackish waters, the heterotrophic dinoflagellate Pfiesteria was the
most abundant eukaryote on PE-MPsand the second most abundant species
on PS (Kettner et al., 2019). Some diatom species have been found to in-
habit the sunlit portion of the plastisphere in marine environments
(Amaral-Zettler et al., 2020). Over time, the abundance of diatoms de-
creases, suggesting that they are replaced by other microorganisms as the
community matures. Eukaryotic microalgae have also been found
within the biofilm on plastic particles (Yokota et al., 2017). The primary
producers found on PE- and PS-MPs in brackish waters were green algae
from the genus Ulva and the class Trebouxiophyceae (Kettner et al.,
2019).
Cyanobacteria are the most studied organisms in the plastisphere,
playing a role in the rate of photosynthesis, the cycling of nutrients, and
other processes in aquatic ecosystems. The filamentous cyanobacteria
Phormidium was the most frequent taxon detected along the surface area
in marine environments (Nava and Leoni, 2021). Additionally, different
species ofcyanobacteria can persist in biofilms, depending on the presence
of dominant species. The photosynthetic filamentous cyanobacteria
Phormidium and Rivularia were found on marine MPs but absent when the
prokaryotic phototroph Prochlorococcus was present in the water (Zettler
et al., 2013). A high density of chlorophyll-a(Chl-a) was observed on
MPs and plastic debris >5 mm in size, suggesting that these plastic particles
are creating net autotrophic “hot spots”in oligotrophic environments
(Bryant et al., 2016).
The size range of NPls is considered too small for bacterial colonization,
even as aggregates (Deschênes and Ells, 2020). Still, NPls can elicit influ-
ence on biofilm formation, particularly at the air-liquid interface. Exposure
to sulfatelatex (hydrophobic) 30 nm PS-NPls resulted in an accumulationin
the strength of floating biofilms of E. coli UTI89 (Zhang and Christopher,
2016). Scanning electron microscopy (SEM) imaging showed that the
30 nm PS-NPls filled in the gaps within the biofilm structure, increasing
the strength of the biofilm. The biofilm communities' interactions with
NPls will be discussed in more detail in the following section.
Further research is needed to confirm whether NPls selectively promote
the growth of certain microorganisms within biofilms in aquatic environ-
ments. As settlement on substrate is dependent on interactions between
the microorganism and substrate (i.e., electrostatic interactions and surface
hydrophobicity), any changes in surface functionalization induced by the
presence of NPls would impact biofilm formation and diversity of the bio-
film communities. In addition, degradation processes will occur in parallel
to the NPls-biofilm interaction. The by-products from the degradation of
NPls will have a pronounced impact on biofilm structure and diversity
(Romera-Castillo et al., 2018). However, the influence of NPl degradation
on biofilm communities is not known. NPl degradation may stimulate mi-
crobial activity (i.e., the release of DOC), thereby promoting or maintaining
a diverse biofilm community. Or it can inhibit microbial activity due to the
release of sorbed pollutants and the production of free radicals (Zhang
et al., 2022). Thus, the tolerance of NPl degradation would depend on
the composition of the biofilm community (i.e., a biofilm community
consisting of species capable of degrading NPls would be able to main-
tain their structure better compared to a biofilm community without
such members).
3.4. Impacts of nanoplastics on aquatic biota
There is a strong relationship between the size of the plastic particles
and their toxicological effects on and interactions with pollutants from
the surrounding environment (Strungaru e t al., 2019). We present a synthe-
sis of the studies of the ecotoxicological impacts of NPls on aquatic biota in
Table 1. The smaller the size of the plastic, the more toxic the impact will be
for the organism (Browne et al., 2008). With NPls, their small size enables
them to sorb contaminants, due to their high surface-to-volume ratio, and
pass through biological barriers and bioaccumulates. In addition to
the toxicity associated with the accumulation of NPls, inherent, sorbed
contaminants such as polybrominated diphenyl ethers (PBDEs) and or-
ganophosphate flame retardants (OPFRs) can be leached out from the
plastics into the environment (Yu et al., 2019). These released co-
leachates can then inhibit the growth of microorganisms (Zhu et al.,
2020). Polycyclic aromatic hydrocarbon (PAH)-sorbed NPls (44 nm)
affected embryonic mitochondrial coupling efficiency and larval mito-
chondrial spare capacity in zebrafish (Trevisan et al., 2020). This “Trojan
horse”effect has the potential to amplify the toxicity of NPls in the long
term (Trevisan et al., 2020;Xu et al., 2020a).
The high surface reactivity of NPls leads to the sorption of NPls on the
surfaces of microorganisms like algae. In the single-celled Chlorella and
multi-celled Scenedesmus species, NPls physically blocked algal pores with
their sorption to the cell surface (Bhattacharya et al., 2010). The blockage
by NPls prevents the gas exchange and light penetration necessary for pho-
tosynthesis and other cellular activities. As a result, algal growth, the rateof
photosynthesis, and antioxidative metabolism are inhibited, as the microor-
ganisms experience oxidative stress (Gao et al., 2021). In the study by
Besseling et al. (2014), a reduction in population growth and Chl-aconcen-
tration in the algae Scenedesmus obliquus was reported under a high concen-
tration (1 g/L) of 70 nm carboxyl NPls (Besseling et al., 2014) (Row 5 in
Table 1). Thedecline in algal growth of Dunaliella tertiolecta was also related
to the observed adsorption of PS-NPls on microalgal surfaces and the de-
crease in algal photosynthesis (Bergami et al., 2017)(Row10inTable 1).
The presence of NPls induces the production of reactive oxygen spe-
cies (ROS) by disrupting the transfer of electrons to oxygen or through
the oxidative degradation of NPls (Amaral-Zettler et al., 2020;Gao
et al., 2021;Yousif and Haddad, 2013). An excess concentration of
ROS can cause damage to DNA, lipids, and other macromolecules, dam-
aging the cellular structure and functionality (Camini et al., 2017). In
Chlorella vulgaris, exposure to carboxyl PS-NPls resulted in a decline
in Chl-a and a rise in ROS production (Hazeem et al., 2020)(Row3in
Table 1). Still, the potential mechanisms of ROS production due to the
presence of NPls remain unclear. UV radiation formed free radicals on
MPs and NPls, so it is possible that they induced the photodegradation
process (i.e., the activation of the immune system) and may generate
ROS similarly (Yousif and Haddad, 2013). Furthermore, the observed
production and mechanism behind ROS generation may differ among
the plastic polymers due to their unique chemical structures. Hence, fu-
ture studies need to focus on the exact mechanisms of ROS production
bythepresenceofNPls.
Other studies support the observation that surface charges play a signif-
icant role in the potential impacts of NPls on algae. Bhattacharya et al.
(2010) found that negatively charged carboxyl PS-NPls elicited less affinity
to, and thus less impact on, algal cells compared to positively charged
amine-modified PS-NPls due to repulsive electrostatic forces and the
negative polarity of cellulose (Table 1). Similar observations were shown
in Bergami et al. (2017), where the inhibition of algal growth in green
microalga D. tertiolecta was more pronounced in the medium consisting of
amine-modified PS-NPl aggregates (Bergami et al., 2017).
In some toxicological studies, contradictory results were reported. The
exposure of Chlorella pyrenoidosa to 1 mg/L of 600 nm PS-NPls resulted in
no apparent impact on chlorophyll fluorescence, total antioxidant capacity
(T-AOC), ROS, and lipid peroxidation, despite inhibition in algal growth
(Wang et al., 2020c). Similarly, notwithstanding a decrease in algal growth
and the subsequent decline in photosynthetic activity, negligible effects in
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
8
ROS production were observed when the red microalgae Rhodomonas
baltica was exposed to carboxyl-modified 50 nm PMMA-NPls (Gomes
et al., 2020).
Other studies observed an increase in growth rate and photosynthetic
activity despite the occurrence of ROS. In Zheng et al. (2021), the enlarge-
ment of both the aggregation rate and the specificgrowthrateofthe
freshwater cyanobacteria species Microcystis aeruginosa was seen when
exposed to 50 mg/L of 60 nm PS-NPls, despite the inhibition of these pa-
rameters in the first 8 days of exposure. The photosynthetic activity of
M. aeruginosa also fluctuated before increasing in the final days of experi-
mentation, despite the deformation of algal cells and the occurrence of
ROS (Zheng et al., 2021). In another study, exposure of M. aeruginosa
to ~200 nm amine-modified PS-NPls resulted in no apparent impact on
algal growth (Zhang et al., 2018). Excess ROS can cause damage to
DNA, lipids, and other macromolecules, resulting in the loss of cellular
structure and functionality (Camini et al., 2017). To address the rise in
ROS production, organisms have an antioxidant defense system that re-
duces the negative impacts of ROS. But if the constant production and
accumulation of ROS overwhelms the antioxidant defense system, oxi-
dative stress and cellular damage can occur. The presence of deforma-
tion and oxidative damage despite no apparent impact on algal growth
and photosynthetic activity suggests the possibility that NPl toxicity is
species-dependent, but more work is needed to determine the exact im-
pacts of NPls on algae.
With biofilm, the surface functionalization of NPls can play a role in the
amount of biofilm that is formed. Exposure to 200 ppm of 20 nm fluores-
cent carboxyl-modified PS-NPls resulted in a significant increase in biofilm
formation by M. adhearens,M. algicola,C. marina,andO. kriegii (Okshevsky
et al., 2020). For P. inhibins, a decrease in total biomass and biofilm formed
was observed despitethe integration of NPls inthe biofilm. The same study
also reported that amine-modified PS-NPls significantly decreased biofilm
formation by all of the tested species except for O. kriegii (Okshevsky
Table 1
The ecotoxicological impacts of nanoplastics on aquatic biota.
Source of NPls
(Composition and size)
Target organism Impact on biota References
Amidine Latex Modified and
Carboxyl Latex Modified PS-NPls
(20 nm)
Chlorella Sorption of NPls; inhibition of photosynthetic activity; increase in
ROS levels
(Bhattacharya et al., 2010)
Carboxyl Modified PS-NPls
(20, 50, and 500 nm)
Chlorella vulgaris Reduction in cell viability and concentration of Chl-a; increase in
ROS production
(Hazeem et al., 2020)
Red Fluorescence-labeled PS-NPls
(600 nm) (at 1 mg/L PS-NPls)
Chlorella pyrenoidosa Inhibition of growth rate; no apparent impact on chlorophyll
fluorescence, T-AOC, ROS, and MDA levels
(Wang et al., 2020a)
Amidine Latex Modified and
Carboxyl Latex Modified PS-NPls
(20 nm)
Scenedesmus Sorption of NPls; inhibition of photosynthetic activity; significant
increase in ROS levels
(Bhattacharya et al., 2010)
PS-NPls Scenedesmus obliquus Inhibition of growth and reduction in Chl-alevels (Besseling et al., 2014)
Amine Modified PS-NPls
(198 to 203 nm)
Microcystis aeruginosa No deformation of cell morphology or apparent effect on growth;
no statistical difference in Chl-alevels after 96 h (in 5 mg/L) but
was observed at 10 and 20 mg/L PS-NPls
(Zhang et al., 2018)
Amine Modified PS-NPls (60 nm) Microcystis aeruginosa Inhibition of algal growth; increase in aggregation rate and algal
density; varied impact on photosynthetic activity; deformation of
algal cells due to physical damages and production of ROS
(Zheng et al., 2021)
PMMA-NPls (~40 nm) Tetraselmis chuii
Nannochloropsis gaditana
Isochrysis galbana
Significant reduction in growth rate (Venâncio et al., 2019)
Yellow-Green Fluorescent Carboxyl
Modified PS-NPls (40 nm)
Dunaliella tertiolecta Significant reduction in growth rate (Bergami et al., 2017)
Amine Modified PS-NPls (80 nm) Synthetic wastewater with chemical oxygen
demand COD of 5200 mg/L anaerobic sludge
4.57 % total solids and 3.64 % volatile solids
Reduction in methane production; inhibition of acidification and
methanation processes in anaerobic digestion
(Zhang et al., 2020)
Amine Modified PS-NPls (500 nm)
(at 2.5 μg/mL)
Chaetoceros neogracile Little impact on cell growth and cell morphology; insignificant
changes to photosynthetic activity
(Seoane et al., 2019)
PMMA-NPls (~40 nm) Thalassiosira weissflogii Significant reduction in growth rate (Venâncio et al., 2019)
Amine Modified PS-NPls Daphnia magna Higher mortality rate; negative effect on body size; reduced clutch size (Besseling et al., 2014)
Amine Modified PS-NPls (53 nm)
Carboxyl Modified PS-NPls
(26 and 62 nm)
Daphnia magna Reduced survival rate
Reduced survival rate after long-term exposure
(Kelpsiene et al., 2020)
Unlabeled and Fluorescence-labeled
PS-NPls (75 nm)
Daphnia pulex Inhibited growth; reduced number of total offspring and number
of clutches; ROS production; inhibition of antioxidant system
(Liu et al., 2019)
Yellow-Green Fluorescent Carboxyl
Modified PS-NPls (40 nm)
Unlabeled Amine Modified PS-NPls
Artemia franciscana Insignificant impact on brine shrimp
Significant up-regulation of genes involved in the molting of A.
franciscana embryos and larvae; increase in mortality after
long-term exposure (14 days)
(Bergami et al., 2017)
Carboxylate Modified PS-NPls
(100 nm) Red Fluorescent
Mytilus edulis Longer gut retention time (Ward and Kach, 2009)
Amine Modified PS-NPls (50 nm) Mytilus galloprovincialis Stimulation of hemocyte extracellular ROS production; negative
impact on mussel immune defense capacity
(Auguste et al., 2020)
Carboxylate Modified PS-NPls
(100 nm) Red Fluorescent
Crassostrea virginica Longer gut retention time (Ward and Kach, 2009)
Amine Modified And Carboxyl
Modified PS-NPls (50 nm)
Crassostrea gigas Decrease in motile spermatozoa; decrease in velocity of the
spermatozoa; decrease in embryogenesis success
(Tallec et al., 2020)
PS-NPls (24 nm) Carassius carassius Disturbance in lipid metabolism and behavior (Cedervall et al., 2012)
PS-NPls (100 nm) Green
Fluorescent
Oreochromis niloticus Reduction in acetylcholinesterase activity; disturbance of
metabolism; increase in superoxide dismutase (ROS production)
(Ding et al., 2018)
PMMA-NPls (~ 45 nm) Dicentrarchus labrax Increased expression in genes and receptors (i.e., peroxisome
proliferator-activated receptors) relating to lipid metabolism;
decreased levels of esterase and alkaline phosphatases in blood
plasma and skin mucus (respectively)
(Brandts et al., 2018)
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
9
et al., 2020). Therefore, the size, surface charge, and concentration of NPls
can influence biofilm formation and induce toxicological impacts.
NPls can interact with biomolecules such as EPS to form an eco-corona
over the plastic particles (Junaid and Wang, 2021). The formation of the
eco-corona was evident in the study by Fadare et al. (2020).Daphnia
magna experienced a reduction in toxicity after 96 h of exposure to posi-
tively charged amino-modified NPls in the presence of humic and fulvic
acids (Fadare et al., 2020). The authors attributed the observations to sur-
face modifications ofNPls caused by humic and fulvic acids, which thereby
reduced the interaction between NPls and the aquaticorganism and the tox-
icity of NPls. In addition to ionic interactions between NPls and biomole-
cules, hydrophobic interactions can induce the binding of NPls to EPS
(Chen et al., 2018;Lundqvist et al., 2008). A confirmation of the impor-
tance of hydrophobicity was provided in experiments using the marine
alga Chlorella sp.; a higher binding affinity for EPS components was
observed with positively charged amino-modified PS-NPls than with nega-
tively charged carboxyl-modified PS-NPls (Natarajan et al., 2020). Due to
the reduced surface area, the eco-corona formed over the NPls yielded a
lower increase in ROS levels and antioxidant enzyme activity compared
to plain NPls.
Filter feeders can easily ingest NPls as individual particles or as aggre-
gates with other materials (Ward and Kach, 2009). The open circulatory
systems of filter feeders can allow NPls to circulate through the blood-
stream and connective tissues to major organs (Browne et al., 2008). Accu-
mulation of NPls in major organs can yield toxic effects on biomarkers
such as growth and development, energy metabolism, and oxidative stress
(ROS production). Exposure of 1 to 50 μg/mL of amino-functionalized
50 nm PS-NPls to invertebrates phagocytes (hemocytes) of the marine bi-
valve Mytilus galloprovincialis increased extracellular ROS generation and
apoptosis within 1 h of exposure (Lehner et al., 2019). Still, little research
has been carried out on the long-term impacts of filter feeders in aquatic
ecosystems. Bivalves play a significant role in the cycling of nutrients and
dissolved organic matter (van der Schatte Olivier et al., 2020). The ability
to filter large volumes of water, for example, helps bivalves remove excess
quantities of nutrients through their incorporation for shell and tissue
growth. A decline in bivalve populations, however, would result in eutro-
phication and promote the occurrence of HABs.
The bioaccumulation of NPls also poses a threat to consumers. In addi-
tion to consuming microorganisms contaminated with NPls, grazers may
considerNPls themselves asfood, causing the dilution of food consumption.
PS-NPls impacted the body size and reproduction of the zooplankton
Daphnia magna (Besseling et al., 2014). In Daphnia pulex, exposure to
75 nm PS-NPls at various concentrations inhibited the expression of stress
defense genes (antioxidant enzymes and heat shock proteins) (Liu et al.,
2019). The reduction of populations of zooplankton would lessen the pred-
atory pressures on phytoplankton and algae,leading to eutrophication and
potential HABs (Kong and Koelmans, 2019).
For organisms occupying higher trophic levels, the cumulative impacts
of NPls are exacerbated a long the food chain t hrough biomagnification. Ad-
verse effects include early mortality, inflammatory responses, the induction
of ROS production, low feeding activity, disturbances in lipid metabolism,
and oxidative damage in the liver (Cedervall et al., 2012;Strungaru et al.,
2019). For the fathead minnow (Pimephales promelas), exposure to PS-
NPls through ingestion and intraperitoneal injection resulted in the
accumulation of the plastic particles in the liver and head kidney and in
the interference with genes responsible for the synthesis and function of
neutrophils and macrophages (Elizalde-Velázquez et al., 2020). After
14 days of exposure to 100 nm PS-NPls, juvenile large yellow croakers
(Larimichthys crocea) experienced a negative impact on their growth rate
and the activities of the digestive enzymes, lipases, and lysozymes (Gu
et al., 2020). This was accompanied by a significant change in the dominant
bacterial phyla in the gut of the large yellow croakers and by the relative
abundance of potentially pathogenic bacteria Alistipes and Parabacteroides.
Moreover, the bioaccumulation of NPls is not unique to fully developed
organisms. Exposing developing zebrafish (Danio rerio)embryos(6hpost-
fertilization) to 51 nm PS-NPls resulted in the accumulation of PS-NPls in
the yolk sac and the migration of PS-NPls to organs like the liver, heart,
and brain during development (Pitt et al., 2018). In another study by Lee
et al. (2022), ~567 nm PP-NPls stained with rhodamine B isothiocyanate
(RBITC) were exposed to zebrafish embryos 72 h post-fertilization for
24 h. The bioaccumulation of the NPls within the gastrointestinal tract
was observed fluorescently via red fluorescent dots (Lee et al., 2022).
This is of concern, as the likely route of exposure for humans is through
the consumption of seafood or drinking water contaminated with NPls.
The adverse effects observed in aquatic biota are expected to occur in
humans as well.
The main caveat in the ecotoxicological studies involves the use of syn-
thetic NPls instead of natural NPls extracted from aquatic environments
(Table 1). Due to their unique formation, transport, and the transformation
processes present in the environment, natural NPls are more likely to be
non-uniform in shape and charge compared to the uniform, spherical
nature of synthetic NPls created under a controlled setting. Furthermore
additional additives, plasticizers, and antimicrobials (i.e., sodium dodecyl
sulfate (SDS)) are added to synthetic NPls to ensure the homogeneity of
the particles (Piccardo et al., 2020). Thus, there will be an observable differ-
ence in toxicological impacts between synthetic and natural NPls due to
morphology, polarity, and presence of additives/plasticizers.
3.5. Current methods of sampling and characterizing nanoplastics
3.5.1. Sampling methods
For sampling NPls, using nets and sieves would be ineffective due to
their large mesh sizes (Kundu et al., 2021). Therefore, other sampling
methods are needed to effectively retrieve NPls in the field (Table 2).
One technique for retrieving NPls involves the use of continuous flow
centrifugation (CFC) to retain the NPls at high throughput and high centrif-
ugal force. NPls of ~160 nm in diameter were effectively retained from ul-
trapure water and filtered and unfiltered river water, depending on the
pump rate and the quantity of recirculation throughout the system
(Hildebrandt et al., 2020). However, CFC is more dependent on the density
of the particles than on the size and shape of NPls.
Membrane filtration is a common technique used in isolating MPs and
NPls. Depending on the pore size of the membrane, NPls as small as 5 nm
in size can be separated. However, membranes with smaller pore sizes
exhibit slower flow rates and are more prone to clogging (Barbosa et al.,
2020). Moreover, pore sizes in the nanometer range are susceptible to
membrane fouling and damage, potentially introducing plastic contamina-
tion to the sample (Schwaferts et al., 2019). Another limitation to this tech-
nique is that the cost of the membrane increases exponentially as pore size
decreases (Li et al., 2020). Separation via sequential filtration using mem-
branes of various pore sizes (from μm to nm) can be used to avoid clogging
and to conserve the size and morphology of NPls for further analysis
(Barbosa et al., 2020;Li et al., 2020). Sequential filtration was used in sep-
arating MPs and NPls in the leachate after plastic tea bags were steeped in
hot water using a 2.5 μm cellulose filter (Hernandez et al., 2019).
Ultrafiltration (UF) allows for the preconcentration, separation, and pu-
rification of NPls from samples through the use of a nano-sized membrane
and hydrostatic force (Li et al., 2020). After UF, the particles are retained
within a low solvent volume. The low volume of solvent easily facilitates
the collection of the particles and reduces any instances of particle aggrega-
tion or degradation (Li et al., 2020). A variation of UF, known as cross-flow
UF, can process large volumes of water through porous sleeves to concen-
trate small particles (Nguyen et al., 2019;Schwaferts et al., 2019). UF
was used to examine the efficiency of removing PS-NPls of varying sizes
(50, 100, and 500 nm) with a highly porous electrospun polyacrylonitrile
(PAN) membrane in a layer-by-layer assembly (Wang et al., 2020b). The
PAN nanofibrous membrane was able to retain most of the 500 nm PS-
NPls but did not effectively retain the smaller PS-NPls (50 and 100 nm).
Thus, UF needs further optimization such as restoring the mechanical
strength ofthe membrane after each filtration cycle and changing the polar-
ity and pore size of the membrane for additional selectivity.
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
10
Field-flow fractionation (FFF) is a separation technique that applies
a perpendicular force on flowing particles to retain them within the
flow channel, depending on their characteristics (i.e., density and size)
(Schwaferts et al., 2019). FFF can separate particles from 1 nm to 100 μm,
depending on the mode and keep the original conditions of the sample
(Laborda et al., 2016). Yet, the presence of contaminants such as NOM
and other types of polymers can hinder the separation and detection of
NPls (Fu et al., 2020). The most universal form of FFF, asymmetric flow
field-flow fractionation (AF4) can be coupled with other techniques
(i.e., py-GC–MS) for the chemical characterization of NPls (Laborda et al.,
2016;Schwaferts et al., 2019). AF4 was able to separate 100 nm fluorescent
PS-NPls after enzymatic digestion in fish when coupled with multi-angle
light scattering (MALS) (Correia and Loeschner, 2018). However, this
method failed to detect PE-NPls because of the high light scattering back-
ground. Some limitations to FFF include uncontrolled particle-membrane
interactions impacting elution time, limited concentration ranges,
and low retention (Laborda et al., 2016;Schwaferts et al., 2019). Further
multistep parameter optimization is required to address these limitations
(Li et al., 2020).
3.5.2. Visualization methods
Scanning electron microscopy (SEM) and energy-dispersive X-ray
spectroscopy (EDS) can be used to characterize the surface morphology
and elemental analysis of NPls (Mariano et al., 2021;Rocha-Santos and
Duarte, 2015). Sample preparation involves first drying and depositing
the sample onto a surface for visualization (Fu et al., 2020). Since plastics
have low conductivity, a thin conductive layer of graphite or metal coats
the sample to reduce any surface-charging effects (Hernandez et al.,
2017;Jakubowicz et al., 2021;Mariano e t al., 2021). Furthermore,the sam-
ple needs to be electrically grounded to prevent any buildup of electrostatic
charge (Table 2). Thus, sample preparation is one of many drawbacks to
SEM; other drawbacksinclude high costs and the destruction of the sample
(Nguyen et al., 2019).
Transmission electron microscopy (TEM) provides information regard-
ing the interior of the sample but is not effective at visualizing NPls due
to their amorphous structure (Jakubowicz et al., 2021;Mariano et al.,
2021). Still, characterization of crystalline and amorphous structures is pos-
sible through selected-area electron diffraction or convergent-beam elec-
tron diffraction (Laborda et al., 2016). The low density and structure of
Table 2
Current methods for the sampling and characterization of NPls.
Method/technique Particle size range Advantages Disadvantages References
Sampling methods
Continuous flow
centrifugation (CFC)
>160 nm Able to retain NPls at a high throughput
and centrifugal force; Can work for
multiple days and large volumes; No
membrane blockage
High equipment costs; dependent on the
density of the particles; Need high
centrifugal force to sample low NPl
concentrations
(Barbosa et al., 2020;Hildebrandt et al.,
2020)
Membrane filtration NPls >100 nm in size
(depending on pore
size –i.e., 100 nm,
450 nm, and 1 μm)
High efficiency; Conserves the size and
shape of NPls
Time-consuming and tedious; high costs;
prone to clogging; membrane fouling for
nano-sized membranes; low flow rates with
small pores; Small volumes
(Barbosa et al., 2020;Hernandez et al.,
2017;Li et al., 2020;Schwaferts et al.,
2019)
Ultrafiltration (UF) 5–50 nm Ability to sample large volume of water;
Little sample degradation and membrane
fouling
Needs further optimization to restore
mechanical strength of the membrane after
each filtration cycle; potential to interact
with the membrane surface
(Laborda et al., 2016;Li et al., 2020;
Schwaferts et al., 2019;Wang et al.,
2020b)
Field-flow fractionation
(FFF) and Asymmetric
flow field-flow
fractionation (AF4)
1nm–100 μm Non-destructive; Keeps original structure
of sample; Does not require a stationary
phase; Fast (<30 min)
Difficult to operate; Time consuming; Steric
inversion; Presence of contaminants hinders
separation; Uncontrolled particle-membrane
interactions
(Barbosa et al., 2020;Fu et al., 2020;
Laborda et al., 2016;Schwaferts et al.,
2019)
Characterization methods
Scanning electron
microscopy (SEM)
30 nm–2μm Enable visual differentiation of particles Sample preparation; high costs; destruction
of the sample; charging effects
(Jakubowicz et al., 2021;Mariano et al.,
2021;Nguyen et al., 2019;Schwaferts
et al., 2019)
Transmission electron
microscopy (TEM)
<1 nm Use detectors to acquire information on
the elemental composition
Time-consuming; Low-throughput; Detection
hampered by the low density of NPls; Use of
heavy-metal stains
(Mariano et al., 2021;Nguyen et al.,
2019;Rocha-Santos and Duarte, 2015;
Schwaferts et al., 2019)
Atomic force microscopy
(AFM)
>0.3 nm High resolution of the surface morphology
of non-conductive samples of a few nm;
Non-destructive; No radiation damage;
Produces 3D images of the sample
Slow; Small area; Not immune to impurities;
Requires contact with the sample
(Mariano et al., 2021;Schwaferts et al.,
2019)
Chemical analysis techniques
Fourier-transform
infrared spectroscopy
(FTIR)/micro-FTIR
1.5–10 μm Reduced analysis time; detecting polar
groups; Use of focal plane array enables
analysis of heterogeneous NPls and fast
measurements
Not applicable for NPls; Dry sample as thin films
or grounded powders; Requires a minimum
sample thickness of ~150 nm; Deposition of
sample onto IR-transparent substrate
(Fu et al., 2020;Jakubowicz et al.,
2021;Mallikarjunachari and Ghosh,
2016;Mariano et al., 2021;Nguyen
et al., 2019;Veerasingam et al., 2020)
Raman spectroscopy 100 nm to ~500 nm Simple sample preparation; No minimum
sample thickness; No destruction of
sample; Low interference from water
Low sensitivity in general; Diffraction
limited; broad auto-fluorescence
backgrounds; Small scattering cross-section;
Sensitive to impurities
(Fang et al., 2020;Ghosal et al., 2018;
Gillibert et al., 2019;Kumar et al.,
2019;Mariano et al., 2021;Schwaferts
et al., 2020;Sobhani et al., 2020)
Optical tweezers Coupled with Raman:
50 nm to sub-20 μm
High throughput chemical identification;
Able to trap and detect aggregates;
Optimization of detecting smaller-sized
and/or individual NPls
Surface-enhanced
Raman spectroscopy
(SERS)
>~50 nm Low water interference; High sensitivity
and unique molecular specificity; Able to
measure Raman scattering from a single
molecule
Lacks reliability and consistency due to
issues with producing consistent and uniform
hotspots for signal enhancement
(Hu et al., 2022;Lv et al., 2020;Xu
et al., 2020b;Zhou et al., 2021)
Tip-enhanced Raman
spectroscopy (TERS)
20 nm High sensitivity; Sample depth of a few nm Poor yield and short lifespan of TER probes (Ivleva, 2021;Kurouski et al., 2020;
Verma, 2017;Yeo et al., 2009)
Pyrolysis coupled with gas
chromatography–mass
spectrometry
(py-GC–MS)
Small masses of NPls
(up to ~50 μg)
Easy to use and run; Little sample
preparation
Dry sample required; Destruction of the
sample; Sensitive to impurities; Unsuitable
for plastic polymers with polar subunits; Poor
reproducibility
(Jakubowicz et al., 2021;Nguyen et al.,
2019;Rocha-Santos and Duarte, 2015;
Schwaferts et al., 2019)
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
11
NPls may also hinder detection by TEM (Table 2). Since NPls are not
electrodense, heavy-metal stains are mandatory for analysis (Nguyen
et al., 2019); however, the stains can impact the chemical structure of the
plastic polymers (Mariano et al., 2021). Other limitations include the
time-consuming process, the low-throughput technique, and the required
high number of particles to be counted to get statistically significant results
(Laborda et al., 2016;Rocha-Santos and Duarte, 2015).
Fluorescence microscopy produces an image of the sample by collecting
fluorescent emissions released by the excitation of fluorophores (Mariano
et al., 2021). Fluorescence microscopy is useful for visualizing plastics be-
cause of the innate ability of plastics to emit fluorescence. Nile Red, a
fluorescent dye, can be added for rapid detection and quantification of
NPls due to its selectivity to plastic particles (Li et al., 2018). However,
Nile Red also stains NOM, so pre-purification is required for environmental
samples.Additionally, the presenceof chemical additives and contaminants
can influence and interfere with the fluorescent signals (Fu et al., 2020).
Therefore, proper pre-treatment of the sample, such as acidic or enzymatic
digestion, is needed to minimize these impurities. Still, often only surface
impurities can be removed. The best spatial resolution with fluorescence
microscopy and Nile Red is 20 μm, making it not suitable for visualizing
NPls (Erni-Cassola et al., 2017). A drawback of the fluorescence micro-
scopic technique is the inability to provide information about the chemical
composition of the samples and the limitation in spatial resolution. Using
stimulated emission depletion (STED) microscopy, a form of fluorescence
microscopy, Nguyen and Tufenkji (2022) visualized NPls of different sizes
and compositions within a model nematode worm Caenorhabditis elegans
(Nguyen and Tufenkji, 2022). However, STED suffers from high costs asso-
ciated with the dyes and the need to test the technique with other plastic
polymers (Nguyen and Tufenkji, 2022).
Atomic force microscopy (AFM) can produce an image of the surface
morphology of non-conductive materials at a resolution of a few nm, mak-
ing it ideal for characterizing NPls (Khulbe and Matsuura, 2000;Mariano
et al., 2021). The best spatial resolution for AFM is 0.3 nm (Table 2).
Other advantages include simple sample preparation (such as spin coating)
and the ability to discriminate between material types on the surface of
polymer blends (Fu et al., 2020;Mariano et al., 2021). Some limitations
of AFM include the fact that it is not immune to outside factors such as
contamination and that due to the required contact with the sample
(Table 2), fragments may be released and produce an incorrect image
(Mariano et al., 2021).
3.5.3. Analytical methods for composition analysis
Infrared (IR) spectroscopy measures the transitions between molec-
ular vibrational energy levels when the dipole moments of the molecule
change by the absorption of infrared radiation (Lee and Chae, 2021).
FTIR produces a spectrum corresponding to specific chemical bonds
based on the adsorption or emissions of the sample. For plastics, the re-
sulting IR spectrum of the unknown sample can then be compared with
spectra of known plastic polymers to identify the properties of the sam-
ple (Mariano et al., 2021). FTIR works well in detecting aliphatic com-
pounds within polar groups (Lee and Chae, 2021). When equipped
with focal plane array (FRA) detectors, the analysis of heterogeneous
NPls within the collected samples can be conducted (Table 2). The
FRA detectors also allow for the fast acquisition of several spectra
within an area through one measurement (Veerasingam et al., 2020).
Micro-FTIR performs visualization and surface chemical mapping of the
sample (Mariano et al., 2021;Rocha-Santos and Duarte, 2015). Micro-FTIR
is best suited for particles larger than ~10 μm, though FTIR may analyze
agglomerates or films of smaller particles (i.e., NPls) (Nguyen et al.,
2019). For example, FTIR was used to analyze NPls found in filtrate-
containing particles, including aggregates, <100 nm in diameter
(Hernandez et al., 2017). The analysis had been performed by placing
3 mL of the dried sample over aluminum foil and then heating it
to form a thin powder film. However, the best spatial resolution with
conventional FTIR is 1.5 μmto10μm. This spatial resolution limits
the analysis of nano-sized particles (Kurouski et al., 2020).
FTIR and micro-FTIR can substrate water vapor and carbon dioxide
spectra from the investigated spectrum (Frias et al., 2010;Rocha-Santos
and Duarte, 2015). The disadvantages of FTIR are that it requires dry
samples to be prepared as thin films or grounded powders before analy-
sis and for the deposition of the sample onto an IR-transparent substrate
(Fu et al., 2020;Nguyen et al., 2019). Optimized minimum sample
thickness is necessary since if the sample thickness is too low, the spec-
tra would be affected by interface interference, substrate effect, and
confinement issues from analyzing particles at a nano-size scale
(Mallikarjunachari and Ghosh, 2016). A promising technique is nano-
FTIR spectroscopy, as it is capable of detecting the subsurface properties
of nano-sized particles but still requires more studies to determine the
origins of the observed peaks in the spectra (Mester et al., 2020).
Another promising technique is Raman spectroscopy, a non-destructive
technique with minimum sample preparation and a highly specificfinger-
print spectrum that is similar but simpler than IR (Sobhani et al., 2019).
This technique allows for the direct analysis of aqueous samples. Raman
spectroscopy is a photon scattering technique; when a sample is exposed
to monochromatic light, the radiation causes some molecules to scatter
the light (Lee and Chae, 2021). This scattering of light contains photons
with shifted energy levels, which can then yield information about molecu-
lar vibrations of chemical compounds of the sample (Käppler et al., 2016).
TheresultingRamanspectrumisthencomparedtoreferencespectraof
defined plastic polymers to identify the chemical composition of the
sample (Fig. 6a and b).
Raman spectroscopy works well in detecting non-polar, symmetric
bonds (abundant in PS and PE) and yields a lower response to polargroups
(Lee and Chae, 2021). The ability to control the diameter of focus of the
incident laser beam (<800 nm) enables Raman to detect a wider spatial
resolution and wavenumber range (0–4000 cm
−1
)(Sobhani et al., 2019).
Raman spectroscopy delivers a specificfingerprint spectrum and exhibits
low interference from water (Fang et al., 2020). Other advantages of
Raman spectroscopyinclude relatively simple sample preparation, no min-
imum sample thickness, and the non-destructive nature of the technique
(Mariano et al., 2021). Raman spectra imaging has been used to identify
NPls down to 100 nm in size (Sobhani et al., 2020). However, the working
spatial resolution of conventional Raman microscopies is 250 nm to
~500 nm due to its low sensitivity and diffraction limitation (Kumar
et al., 2019;Kurouski et al., 2020).
Raman spectra are often accompanied by auto-fluorescence back-
grounds, which tend to overshadow the less intense peaks and complicate
the identification process (Ghosal et al., 2018). The sample's chemical com-
position may be identified via a fluorescence removal algorithm or the use
of cellulose filters, both of which result in a lower fluorescence background
at an increased cost (Barbosa et al., 2020). A significant limitation of Raman
spectroscopy is its small scattering cross-section, which leads to a weak sig-
nal (Sobhani et al., 2019). Furthermore,the lateral resolution of the Raman
signal is limited to the diffraction limit of the laser spot needed to excite the
Raman emission (Fang et al., 2020). The analysis of smaller-sized NPls
would result in a lot of false-positive and false-negative signals within the
Raman spectrum (Table 2). The Raman signal is also sensitive to the pres-
ence of color, fluorescence, contaminants, and additives within NPls
(Fu et al., 2020). Pixel resolution improvement through analytical algo-
rithms and color offsetting can enhance the weak Raman signal and bet-
ter identify and visualize NPls (Fang et al., 2020).
Optical tweezers can trap and manipulate MPs and NPls dispersed in
aqueous bodies. When coupled with Raman spectroscopy, optical tweezers
are able to chemically identify NPls between sub-20 μm and 50 nm
(Gillibert et al., 2019). Raman tweezers were able to trap and detect aggre-
gates of NPls, with diameters ranging between 50 nm and 90 nm, and indi-
vidual NPls of diameters of 300 nm and larger (Gillibert et al., 2019). Still,
more work is needed to detect smaller-sized and/or individual NPls.
With SERS, single molecules of NPls can be measured when situated
within SERS “hotspots”induced by the presence of metallic NPs (Hu
et al., 2022;Lv et al., 2020;Xu et al., 2020b). The chemical information
of ~50 nm PS-NPls was obtained when surrounded by silver NPs through
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
12
SERS mapping (Zhou et al., 2021). However, inhomogeneous hotspot
densities resulted in a non-uniform enhancement effect on the Raman
signals. Furthermore, SERS suffers from the diffraction limit of light
and the inability to provide information on morphology (Verma,
2017;Zhou et al., 2021).
TERS can detect NPls by using a metallic nanotip to enhance the Raman
signals one paper that reported the use of TERS for analyzing nano-sized
plastics; a thin film made up of PS and polyisoprene (PI) was examined
with a spatial resolution of ~20 nm (Yeo et al., 2009). Thus, future applica-
tion of TERS is needed. One drawback of TERS is the poor yield and short
lifespan of the TER probes, as the metallic probes can undergo oxidation
when in contact with air for a long period (Kurouski et al., 2020).
Some destructive techniques include py-GC–MS (Table 2). This multi-
step method involves heating the sample, separating the gaseous products
via gas chromatography, and analyzing the sample using mass spectrome-
try (Nguyen et al., 2019). Py-GC–MS was applied to analyze NPls in the
a)
b)
Fig. 6. a) Raman spectra collected from different plastic polymers sampled from the waters of San Francisco, USA - retrieved from SLoPP-E spectra library (Munno et al.,
2020); b) The functional groups of each plastic polymer with their respective characteristic peaks (Nava et al., 2021).
N. Kokilathasan, M. Dittrich Science of the Total Environment 849 (2022) 157852
13
North Atlantic Subtropical Gyre (Ter Halle et al., 2017). Another study to
detect NPls used py-GC–MS for surface and groundwater samples following
ultrafiltration and hydrogen peroxide digestion (Xu et al., 2022). The ad-
vantage of this technique is that analysiscan be done in a single run without
using solvents, thereby avoiding background noises (Rocha-Santos and
Duarte, 2015). This technique can also identify small masses of NPls
(to ~50 μg), but this sensitivity may be a barrier if the NPls are attached
to or embedded in other materials (Nguyen et al., 2019). There are sev-
eral shortcomings of this technique, including the destructive aspect of
the method, the poor reproducibility, the inability to determine the
shape and size of the NPls, and the fact that each pyrolysis type has a
specific temperature behavior (Li et al., 2020;Rocha-Santos and Duarte,
2015;Xu et al., 2022). Thus, more work is needed to improve the limit
of detection of py-GC–MS and its sensitivity to impurities.
In summary, there are several limitations to the current techniques for
the detection and characterization of NPls. The spatial resolution does not
allow the detection and analysis of individual NPls and/or NPls smaller
than 100 nm in size. Therefore, a thorough investigation on the possible
set of protocols for the detection of NPls in aquatic environments should
be the focus of future research. The investigation can be done either
by improving on existing methods (i.e., the optimization of nano-FTIR
and optical tweezers) or developing new techniques capable of separat-
ing, visualizing, and characterizing NPls in environmental samples. The
establishment of consistent, reliable, and standardized sets of proce-
dures will aid in further research ontheformationanddegradationof
NPls and the impacts of NPls on aquatic biota.
4. Future work and conclusion
As plastic production continues to rise, more and more plastic debris are
transported to aquatic environments, where they are relatively persistent.
Under unique environmental factors (i.e., salinity, NOM, and microbial
colonization), plastics can undergo transformation (i.e., biofilm forma-
tion), degradation, and transport processes that can influence the rate
of NPl formation. Thus, it is urgent to study the by-products of plastic
degradation at the nano-scale. This review highlights the gaps and
achievements in our knowledge about the potential factors affecting
the formation and degradation of NPls, the ecotoxicological impacts of
NPls on aquatic biota, and the detection and analysis of NPls in aquatic
environments.
One proposed focus for further investigations into the formation of NPls
from different plastic polymers is performing experiments on the exposure
of manufactured MPs and NPls using typical fate and transport processes in
aquatic environments. By observing and analyzing the resulting NPls, the
lifespan and degradation rate of NPls can be determined. Additionally,
the use of irregular NPls over synthetic, spherical NPls in ecotoxicological
studies can serve as a reliable representative of the actual toxicity of NPls
in aquatic biota. Ideally, irregular-shaped NPls are to be retrieved from
field studies or produced via exposing synthetic MPs and NPls to envi-
ronmental conditions. The previous future directions of research are de-
pendentonasetofdefined protocols capable of characterizing NPls in
aquatic systems. Developing new analytical methods and improving
existing ones are necessary to address the remaining gaps in knowledge
regarding the composition and fate of NPls. For sampling, this means
creating unified protocols that target the isolation of NPls and remove
any contaminants through pre-treatment procedures for further analy-
sis. For the composition determination of NPls, the amendment of
such current techniques or novel techniques as nano-FTIR, TERS, and
optical tweezers in Raman microscopy are promising tools for future re-
search on NPls.
NPls are an emerging threat to aquatic biota and humans in ways that
MPs and macroplastics are not. Therefore, NPl pollution should be priori-
tized in the field of plastic research. We identified future directions of re-
search that can address the major gaps in knowledge regarding NPls. A
major step would be to create a set of protocolsusing novel or existing tech-
niques to sample and characterize NPls from aquatic systems.
CRediT authorship contribution statement
NK and MD: Conceptualization; NK: Data curation and Formal analysis;
MD: Funding acquisition;
NK and MD: Investigation; Methodology; MD: Supervision; NK:
Validation; Visualization; Roles/.
NK and MD: Writing –original draft; NK and MD Writing –review &
editing.
Declaration of competing interest
The authors declare that they have no known competing financial inter-
ests or personal relationships that could have appeared to influence the
work reported in this paper.
Acknowledgments
This project was funded by the National Sciences and Engineering
Research Council of Canada (NSERC Discovery Grant), NSERC Alliance
“Source-specificidentification, characterization and control of
microplastics across a remote, rural and urban”, and the Canada Foun-
dation for Innovation and Ontario Ministry of Research and Innovation
(Leaders Opportunity Fund, Grant Number 22404) to MD. We thank our
editor Jennifer Krisillas for careful proofreading English and the
reviewers for their thorough reading of our manuscript and helpful
comments.
Appendix A. Supplementary data
Supplementary data to this article can be found online at https://doi.
org/10.1016/j.scitotenv.2022.157852.
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