Nanoplastics : state of knowledge and environmental and human health impacts.
Date: 2023-02-14
Institutions: Directorate-General for Environment
Subjects: circular economy, environmental impact, environmental protection, foresight, health risk, plastic waste, plastics, plastics industry, pollution control measures

Science for Environment Policy

FUTURE BRIEF:
Nanoplastics: state of knowledge 
and environmental and human 
health impacts

February 2023
Issue 27

Environment

2

Science for Environment Policy
Nanoplastics: state of knowledge and environmental and human health impacts

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Acknowledgements

We wish to thank the scientific advisor, Professor Anders Baun (Department of 
Environmental Engineering, Technical University of Denmark), for his input to this 
report. We would like to thank the European Commission’s Joint Research Centre for its 
comments and contribution to this report, in particular: Susanne Belz, Susanne Bremer-
Printed by the Publications Office of the European Union in Luxembourg
Hoffmann, Douglas Gilliland and Birgit Sokull-Kluettgen.
In addition, we wish to thank Andrej Kobe, European Commission, for his input to this report.

The contents and views included in Science for Environment 
Policy are based on independent research and do not necessarily 
reflect the position of the European Commission.

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To cite this publication:

Science for Environment Policy 
(2023) Nanoplastics: state of 
knowledge and environmental and 
human health impacts. Future Brief 
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Science for Environment Policy 
Future Brief 27:
Nanoplastics: state of knowledge  
and environmental and human  
health impacts

4

Contents

1. 

Introduction: nanoplastics in context 

1.1 What we do – and don’t – know 

1.2 Regulatory landscape  

2. Detection and assessment  

2.1 Methods of characterisation and quantification 

2.2 Considerations for policy 

3. Nanoplastics in the environment 

3.1 Sources and routes into the environment 

3.2 Degradation 

3.3 Environmental fate 

3.4 Estimating nanoplastics in field samples 

3.5 Nanoplastics in aquatic organisms 

3.6 Terrestrial habitats 

4. Ecotoxicity of nanoplastics 

4.1 Potential mechanisms of toxicity  

4.2 Effect concentrations 

4.3 Species sensitivity distributions and preliminary threshold concentrations for nanoplastics 

4.4 Eco-corona 

4.5 Electrostatic charge 

4.6 Particle and chemical toxicity in plants 

4.7 Chemical additives in nanoplastics 

5. Impacts on human health 

5.1 Routes into the human body 

5.2 Toxicity in humans 

5.3 Risk assessment 

5.4 Biomonitoring of plastic-associated chemicals and metabolites 

5.5 Nanoplastics in the food chain 

6. Summary 

References  

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6

1. 

Introduction: nanoplastics in context

Plastics are highly versatile and are ubiquitous in 
society, fulfilling a wide array of valuable functions 
in  our  economy  and  daily  lives.  Their  adaptability 
has led to them being used in consumer products 
such  as  cosmetics,  cooking  utensils  and  food 
packaging,  among  others,  as  well  as  in  medical 
devices and construction. While there has been a 
strong push to ban single-use plastics and develop 
plastic-free and recyclable packaging and products 
in recent years, notably as part of the EU’s Plastics 
Strategy11  and  Circular  Economy  Action  Plan22, 
plastic disposal remains a key threat to our natural 
environment,  and  the  material  is  accumulating 
in  our  soil  and  seas  in  unprecedented  amounts. 
It  has  been  suggested  that  an  extra  33  billion 
tonnes  of  plastic  will  be  added  to  the  planet  by 
2050 (Galloway, 2015) and some have suggested 
the  current  era  may  even  be  referred  to  as  the 
‘Plasticene’ (Reed, 2015).

Overcoming the challenge of plastic pollution has 
been named “the defining challenge of our times” 
by  the  United  Nations  Environment  Programme 
(UNEP,  2018).  The  impact  on  our  environment 
is  growing  to  such  a  scale  that  plastic  litter  is 
considered  by  some  as  an  emerging  threat  to 

how our planet functions and stabilises itself at a 
global scale (Galloway and Lewis, 2016). Scientists 
have even proposed marine plastic pollution levels 
be included as one of the ‘planetary boundaries’ – 
a set of nine environmental thresholds that must 
not  be  exceeded  in  order  to  protect  and  support 
the  healthy  function  and  development  of  our 
planet  and  its  communities  (Steffen  et  al.  2015). 
Plastic pollution does indeed fulfil two of the three 
boundary criteria in that it is likely irreversible and 
globally  ubiquitous.  More  knowledge  is  required 
to  ascertain  if  it  also  fulfils  the  third:  disruption 
to ecosystems or, on a wider scale, the entirety of 
the  Earth  system  (Villarrubia-Gómez,  Cornell  and 
Fabres,  2018).  Some  researchers  suggest  that 
since  plastic  pollution  is  occurring  faster  than  it 
is  possible  to  conduct  safety  assessments  and 
monitoring,  it  is  exceeding  the  ‘safe  operating 
space’ of the planetary boundary for novel entities, 
demanding  prompt  action  (Persson  et  al.,  2022). 
The need for such action is especially pressing in 
light of the potentially synergistic effects of other 
types of chemical pollution; low capacity of some 
countries,  globally,  to  abate  it  through  regulation 
and  enforcement;  and  persistent  threat  posed  by 
plastics already released.

Of the 6.3 billion tonnes of plastic waste generated by 2015, nearly 80% 
has ended up in either landfill or the natural environment (and only 9% 
has been recycled) 

– 

termed 

sometimes 

Microplastics  may  be  deliberately  manufactured 
materials 
primary 
microplastics – but there are very few intentionally 
produced plastic nanomaterials. In the environment, 
plastic also breaks down to form small fragments 
known  as  incidental  or  secondary  microplastics 
and,  at  even  smaller  scales,  nanoplastics. 
Research has focused on microplastics rather than 
nanoplastics  due  to  their  greater  visibility  and 
availability of suitable detection methods. There is 
growing evidence of human and wildlife exposure 
to  microplastics  in  the  environment,  and  some 

Geyer, Jambeck and Law, 2017  

data on the potential hazard to ecosystem health, 
but their impacts on human health remain unclear. 

For  nanoplastics,  the  situation  is  even  less  clear, 
and  data  even  more  sparse.  While  debate  is  still 
ongoing about the size definitions for microplastics 
(Hartmann  et  al.,  2019),  it  is  an  open  question 
at  which  size  a  microplastic  particle  becomes  a 
nanoplastic  particle.  This  discussion  is  more  than 
a  formality.  Due  to  their  smaller  size,  concern  is 
rising  over  the  potential  impacts  of  nanoplastics 
specifically on the environment, human food chain, 

1 
2 

https://environment.ec.europa.eu/strategy/plastics-strategy_en
https://environment.ec.europa.eu/strategy/circular-economy-action-plan_en

Plastics are highly versatile and are ubiquitous in 

society, fulfilling a wide array of valuable functions 

in  our  economy  and  daily  lives.  Their  adaptability 

has led to them being used in consumer products 

such  as  cosmetics,  cooking  utensils  and  food 

packaging,  among  others,  as  well  as  in  medical 

devices and construction. While there has been a 

strong push to ban single-use plastics and develop 

plastic-free and recyclable packaging and products 

in recent years, notably as part of the EU’s Plastics 

Strategy11  and  Circular  Economy  Action  Plan22, 

plastic disposal remains a key threat to our natural 

environment,  and  the  material  is  accumulating 

in  our  soil  and  seas  in  unprecedented  amounts. 

It  has  been  suggested  that  an  extra  33  billion 

tonnes  of  plastic  will  be  added  to  the  planet  by 

2050 (Galloway, 2015) and some have suggested 

the  current  era  may  even  be  referred  to  as  the 

‘Plasticene’ (Reed, 2015).

Overcoming the challenge of plastic pollution has 

been named “the defining challenge of our times” 

by  the  United  Nations  Environment  Programme 

(UNEP,  2018).  The  impact  on  our  environment 

is  growing  to  such  a  scale  that  plastic  litter  is 

considered  by  some  as  an  emerging  threat  to 

7

and human health; they are difficult to detect and 
isolate  from  environmental  samples  via  standard 
separation and analytical methods, and there is a 
risk that these tiny particles may be more efficient at 
crossing biological membranes than microplastics. 

This  Future  Brief  elucidates  the  current  situation 
regarding  nanoplastics  and  the  environment, 
from their detection and analysis through to their 
potential  health  risks,  dispersal  pathways,  and 
future outlook. 

Side view of a woman holding microplastics in her hands with composition infographics (©Shutterstock, photo by SIVStockStudio)

8

1.1 What we do – and don’t – know

According  to  the  European  Chemicals  Agency 
(ECHA)  and  Organisation  for  Economic  Co-
operation  and  Development  (ECHA,  2019;  OECD, 
2021), microplastics are small particles of plastic 
typically no more than 5 millimetres (mm) across. 
This  working  definition  for  a  microplastic  aligns 
with  that  proposed  at  a  National  Oceanic  and 
Atmospheric  Administration  (NOAA)  meeting  in 
2008 and, while not formalised as an international 
standard, remains widely used (Arthur, Baker and 
Bamford, 2009; Hartmann et al., 2019). 

Generally, nanoplastics are considered to be the 
plastic particles of the smallest size, i.e. smaller than 
microplastics.  Although  the  EC  Recommendation 
on  the  Definition  of  Nanomaterials3  (including 
plastics)  states  that  they  contain  particles  in  the 
size  range  1-100  nm  (EC,  2022b),  in  scientific 
literature  there  is  currently  no  agreed  distinction 
on  the  size  of  nanoplastics  as  opposed  to 

microplastics,  (Mitrano,  2019;  Hartmann  et  al., 
2019).  For  example,  some  set  the  threshold  at 
1  micrometre  (µm;  0.001  mm);  others  place  a 
nanoplastic’s  upper  size  limit  10  times  lower, 
at  100  nanometres  (nm;  0.0001  mm)  (Ng  et  al., 
2018;  EFSA,  2021).  Definition  of  a  nanoplastic 
as  a  particle  measuring  no  more  than  1μm 
(0.001mm)  across  in  any  one  dimension  is 
widely accepted as a guiding definition (Gigault et 
al.  2018;  Dick  Vethaak  and  Legler,  2021;  SAPEA, 
2019). It is to be noted that this threshold as well 
as the approach (longest dimension) deviate from 
the definitions of nano-objects and nanomaterials, 
which  share  the  prefix  but  are  usually  related  to 
features at 1-100nm scale. In addition, the longest 
dimension of a plastic particle may not be the most 
appropriate  characteristic  in  assessing  risk  –  the 
smallest  dimension  often  indicates  the  likelihood 
of interaction with biological systems.

Fig. 1: Classification of plastic particles by their size and size references, according to definition of nanoplastic 
by Hartmann et al. (2019) © Andreas Mattern/ UFZ

3 

https://ec.europa.eu/environment/chemicals/nanotech/pdf/C_2022_3689_1_EN_ACT_part1_v6.pdf

9

Nanoplastics include conventional, bio-degradable 
and bio-based plastics. Currently, there is no firm 
evidence that the raw materials chosen to create 
a  plastic  product  affect  the  risks  resulting  from 
that  product  entering  our  environment.  Plastic 
polyethylene  terephthalate  (PET)  bottles  derived 
from  petrochemicals  exhibit  the  same  chemical 
and  physical  properties  and  behaviour  in  the 
environment as plastic bottles created from sugar 
cane, for example, and so “bio-based alternatives” 
to  conventional  plastic  particles  “are  a  priori 
included  as  microplastics  [and  nanoplastics]  until 
there is proof for their low potential risk” (EC, 2017). 

The  terms  ‘plastic’  and  ‘polymer’  are  often  used 
interchangeably, but they are not the same thing; 
not  all  polymers  are  plastic.  Polymers  consist  of 
chains of joined molecules or monomers, and may 
be found in natural materials such as wool, cotton 

and wood – and natural polymers are not plastics. 
Plastic is a material produced from synthetic (e.g. 
polypropylene)  or  semi-synthetic  polymers  (e.g. 
viscose),  or  containing  a  high  proportion  of  these 
polymers,  and  often  additives.  Plastics  may  be 
moulded  when  soft  (usually  due  to  heating)  and 
retain  a  given  shape  when  hardened  (either  by 
cooling or chemical processing). Another group of 
polymeric materials, elastomers (elastic polymers, 
commonly  known  as  rubber),  are  often  included 
in  the  term  plastic,  although  they  have  different 
properties.  In  materials  science,  elastomers  are 
not  plastics  because  they  do  not  have  structural 
integrity under a load (they stretch), however when 
we talk of plastics in the environment, it is useful to 
include elastomers as this encompasses particles 
from tyre wear, for example. 

Fig. 2 All plastics contain polymers, but 'plastic’ and ‘polymer’ are not the same thing. 
Source: Aznan, Shutterstock

10

Box 1: Features of nanoplastics

Nanoplastics:

 y are  often  defined  as  particles  measuring  no 
more than 1 µm (0.001 mm) across in any one 
dimension  (but  risk  may  be  related  to  their 
smallest dimension)

 y include  conventional,  biodegradable  and  bio-

based plastics

 y in the context of environmental science, include 

elastomers

 y include primary, manufactured  

nanomaterials, and secondary nano-sized 
particles which are the product of larger  
plastic particle degradation 

1.1.1 How do nanoplastics differ from other plastic particles and nano-sized materials?

Although  nanoplastics  share  many  characteristics  with 
microplastics,  they  are  a  behaviourally  distinct  type 
of  environmental  contaminant.  As  summarised  by 
Brewer, Dror and Berkowitz (2020), nanoplastics are:

 y small  enough  to  penetrate  more  and  different 
biological  barriers,  which  in  turn  increases  their 
potential  for  toxicity  and  transport  of  other 
adsorbed  toxins  within  organisms  –  including 
humans

 y far more difficult to quantify in natural samples, 
leading to significant knowledge gaps when compared 
to the body of research tackling microplastics

 y able  to  adsorb  (hold  onto  in  a  thin  encasing  film) 
significant amounts of other pollutants that are 
known to be harmful (e.g. heavy metals, pesticides, 
pharmaceuticals,  endocrine  disrupting  chemicals, 
microorganisms)  due  to  their  high  surface  area-to-
volume ratio, thereby increasing the potential mobility 
and dispersal of these pollutants. (Microplastics also 
adsorb  substances,  however  nanoplastics  have  an 
even greater capacity to do so as a result of their 
higher specific surface area.) 

 y more reactive than larger plastic particles due 
to their high surface area-to-volume ratio, therefore 
more  likely  to  interact  with  other  materials  in  their 
environment and potentially more likely to destabilise 
or aggregate

 y small enough to avoid plastic removal processes 
such as physical straining through soil or sediment

As  with  all  plastics,  nanoplastics  may  leach  chemical 
additives  and  potentially  present  by-products  and 
monomers, possibly at a faster rate than larger particles 
due  to  their  greater  relative  surface  area.  This  raises 
concerns about the potential toxicological hazards of these 
substances. Whether coatings that form on nanomaterials 
as  they  enter  the  environment  or  organisms  (eco-  or 
protein  coronas)  influence  their  behaviour,  is  an  ongoing 
topic of research.

For  safety  assessment  of  nanomaterials 
(natural, 
manufactured  or  incidental),  regulators  are  concerned 
with a number of issues – which also apply to nanoplastics 
(Allan et al., 2021) – such as:

 y clear definitions

 y ensuring there are appropriate analytical methods for 
detecting them in a variety of matrices (e.g. in goods 
and the environment)

 y appropriate toxicological methods for hazard 

assessment 

indicative  of  potential 

The  level  of  plastic  pollution  in  the  environment 
levels  of  nanoplastic 
– 
pollution  –  makes  such  knowledge  gaps  an  urgent 
regulatory  measures. 
focus  for 

investigation  and 

The extremely large volumes of plastic contaminants entering the 
environment compared to most engineered nanoparticles further 
underscore the necessity of singling out nanoplastics as a unique 
potential hazard 

Brewer, Dror and Berkowitz, 2020

 
11

1.1.2 Primary and secondary nanoplastics

Plastic  particles  enter 
the  environment  via 
multiple pathways, with two key origins: primary 
(nanoplastics that are intentionally manufactured) 
and  secondary  (those  that  are  formed  as  larger 
plastics degrade). Primary nanoplastics are far less 
prevalent  than  primary  microplastics,  and  most 
plastic debris found at the nano-size is secondary 
(Bianco  and  Passananti,  2020).  The  major  source 
of environmental nanoplastics is physical abrasion 
of  larger  plastic  products  across  their  lifecycles 
(Gangadoo et al., 2020; Nanotechnology Industries 
Association, 2020). 

Microbeads  –  small  solid  plastics  that  are  added 
to exfoliating face scrubs and other personal care 
products  –  are  a  well-known  example  of  primary 
plastic particles. Primary nanoplastics are also 
added  to  a  wide  range  of  products  for  diverse 
purposes,  also  being  found  in  personal  care  and 
cosmetics products, in products for biomedical or 
laboratory applications, and textiles (EC, 2017). Table 
1 gives a list of micro- and nanoplastic functions. 
These  particles  are  designed  for  commercial  use, 
and make their way into the environment via routes 
such as wastewater (Bianco and Passananti, 2020).  

Table 1. Functions of micro- and nano-sized plastic particles in different products. The source table (adapted 
from  EC,  2017)  specifically  refers  to  microplastics.  However,  the  upper  limit  of  nanoplastics  is  currently  a 
matter  of  debate  and  many  of  these  functions  also  apply  to  nanoplastics  as  currently  defined.  Many  of  the 
particles referred to here lie in the nanoplastic size regime (below 1 μm (=0.001 mm) in any one dimension). 
Opacifying nanoplastics, for example, can be as small as 170 nm (=0.0002 mm) in size, while plastics added to 
improve paint clarity are generally below 1 μm (=0.001 mm) in size.

Function

Products

Abrasive/exfoliating

Cosmetics, detergents, industrial blasting 
abrasives

Emulsifier, suspending agent

Cosmetics, detergents, paints

Binding

Filler

Cosmetics, paints, inks, concrete

Construction (wall and joint fillers, self levelling 
compounds/screeds)

Controlled release of ingredients

Pharmaceuticals (nano capsules), cosmetics, 
fertilisers, crops, detergents (enzymes)

Film forming

Cosmetics, polishing agents

Surface coating

Paper making, polishing agents

Improved chemical and mechanical resistance

Coatings, paints, floor coatings, polymer 
cement. Coatings can include, for example, 
house paints with water repellency; glass and 
façade coatings for high-rise buildings; for 
transport such as cars, aeroplanes, ships and 
for structures such as bridges; or coatings used 
to protect electronic products – both consumer 
and industrial.

12

Function

Products

Fluid absorbents

Thickening agent

Aesthetics

Flocculant

Dewatering

Nappies, water retainer for farming, agriculture, 
horticulture

Paints, cosmetics, concrete, oilfield use (drilling 
fluids)

Coloured microplastics in make-up, structural 
effects of paints, enhanced gloss level of paints

Waste water treatment, oilfield use, paper 
making

Paper making, dewatering of sewage sludge, 
manure

Dispersing agent

Paints, coatings (pigments)

Opacifying agent

Anti-static agent

Cosmetics

Cosmetics and hair care

Fig. 3. One use of primary nanoplastics is in scratch resistant coatings. Source: Roman Zaiets, Shutterstock

13

Box 2: Intentionally manufactured nanoplastics

Very small quantities of nanoplastics are produced 
as test or reference materials in scientific research. 
For example:

 y Germany’s  Federal 

Institute  for  Materials 
(Bundesanstalt  für 
Research  and  Testing 
Materialforschung und -prüfung (BAM)) intends 
to  produce  small  quantities  of  nanoplastic 
particles  as  reference  materials  to  use  in  its 
work  on  the  EU’s  POLYRISK  and  PlasticsFatE 
projects,  both  of  which  are  studying  the 
effects  of  nanoplastics  on  human  health.  

 y Research  projects  for  detecting  nanoplastics 
in  the  environment  have  used  a  method  to 
synthesise  nanoplastic  particles,  adding  a 
chemically  entrapped  metal  which  acts  as  a 
tracer  (Mitrano  et  al.,  2019).  This  provides  an 
effective way to accurately detect nanoplastics 
in complex media such as sewage sludge.

 y Nanoplastics  coloured  with  fluorescent  dyes 
have  been  used  to  investigate  absorption 
across biological membranes (Catarino, Frutos 
and Henry, 2019).

 y Polystyrene nanoparticles are used to calibrate 

scientific instruments.

nanoplastics,  meanwhile, 

are 
Secondary 
created via degradation and mechanical abrasion 
of  larger  plastic  debris,  including  microplastics, 
and  therefore  tend  not  to  have  a  uniform  shape 
and morphology due to their method of formation. 
They are largely shaped by their initial composition 
and the ageing or degradation processes they have 
been subject to (see section 3.2), exhibiting a wide 
array of physical and chemical properties (Brewer, 
Dror  and  Berkowitz,  2020).  Sources  of  secondary 
nanoplastics  include  tyre  wear,  synthetic  textiles, 
agricultural  plastics,  fishing  equipment  and 
packaging.  For  example,  one  study  found  that  on 
average 1.4 to 2.1 mg of nanoplastic particles was 
released  from  a  single  gram  of  polyester  fleece 
textile  during  washing  and  abrasion  experiments, 
indicating  synthetic  fabrics  may  be  a  significant 
source  of  nanoplastics  (Yang,  Luo  and  Nowack, 
2021). Tyre wear is a well-known source of micro- 
and nanoplastics in the environment, with 0.3% of 
the  21,200  t/y  of  tyre  wear  particles  lost  to  the 
environment  at  the  nanoscale,  i.e.  below  0.1µm 

(Knight  et  al.,  2020;  Prenner  et  al.,  2021).  These 
particles may also derive from mulches made from 
old tyres, used in playground surfaces and sports 
pitches.

A further source of nanoplastic pollution is emerging 
in  the  form  of  their  unintentional  formation  and 
release  to  the  environment  as  a  by-product  of 
other  manufacturing  processes.  One  example  of 
this  is  in  the  rapidly  growing  three-dimensional 
printing sector (Rodríguez-Hernández et al., 2020). 
Following  a  printing  procedure,  equipment  is 
cleaned  with  an  alcohol/resin  mixture,  which  can 
unintentionally form small nanoplastic particles if 
it is exposed to UV radiation (e.g. sunlight). If this 
mixture  is  subsequently  disposed  of  incorrectly  – 
an emerging risk given the rising domestic use of 
such  printing  devices  –  these  nanoplastics  could 
find their way into wastewater and agglomerate in 
seawater,  forming  a  new  and  as  yet  unquantified 
potential source of nanoplastics.

1.1.3 Knowledge gaps

There  are  many  emerging  applications  for  –  and 
sources  of  –  nanoplastics,  but  many  knowledge 
gaps  remain  that  must  be  addressed  in  order  for 
us  to  use  these  materials  safely.  These  require 
advances in the ability to sample, isolate, detect, 
quantify,  and  characterise  nanoplastics,  as 

well  as  standardised  methodologies  to  permit 
reproducible  findings  and  comparable  data.  
For example, there are challenges in the ability to 
reliably detect, quantify, and identify target nano-
scale particles – particularly in complex mixtures. 
This  may  present  challenges  in  enforcement  of 

 
 
14

regulation addressing them (see e.g. ECHA, 2020a; 
discussed  in  Chapter  2).  In  August  of  2022,  the 
COM  draft  proposal  on  intentionally  added  micro 
(and nano) plastics was released, with no lower size 
limit  –  thus  intentionally  added  nanoplastics  will 
be  covered  by  the  regulation  (EC,  2022a).  REACH 
restricts  manufacturers  from  knowingly  adding 
nanoplastics  to  their  products,  however  it  is  also 
acknowledged that at present analytical methods 
are unable to check products for compliance below 
0.1 µm for spherical particles.

Compared  to  our  knowledge  of  marine  plastic 
pollution, there remains a further lack of research 
on  the  behaviour  and  fate  of  plastic  particles 
in 
terrestrial 
environments.  More  work  looking  into  how  their 
morphology  affects  both  behaviour  and  impacts 
is  needed.  It  is  known  that  microplastics  arise  as 

freshwater,  atmospheric  and 

fragments, pellets, filaments, fibres, broken edges, 
beads,  and  irregularly-shaped  particles  (Rosal, 
2021);  are  some  shapes  of  nanoplastics  more 
problematic  than  others?  How  they  interact  with 
pharmaceuticals and other emerging contaminants 
is another important area of research.

Assessing  risks  of  nanoplastics  to  human  health 
requires  data  on  exposure.  When  considering 
nanoplastics  in  the  food  chain,  for  example, 
Toussaint et al. (2019) identify the heterogeneity of 
the methodologies and experimental designs as a 
key challenge. Exposure to micro- and nanoplastics 
via other routes such as inhalation is now receiving 
attention  (Prata,  2018;  also  see  sections  3.4  and 
5.1), but standardised methodology is yet to be 
implemented.

Research into (micro- and) nanoplastics has started only recently  
and thus, there are many unresolved issues of terminology, definitions, 
sampling, characterisation and the assessment of hazard and exposure 
that, in combination, make it difficult, even impossible, to evaluate and 
regulate the potential risks of nanoplastics 

Allan, Sokull-Kluettgen and Patri, 2020 

 
15

Other research needs identified in the European Strategy 
for Plastics in a Circular Economy are being addressed 
in  the  European  Cluster  on  Health  Impacts  of  Micro- 
and  Nanoplastics  (CUSP)12.  CUSP  research  results  will 
contribute  to  the  health-relevant  aims  of  the  plastics 
strategy and the Bioeconomy Strategy, as well as the 
REACH  restrictions  on  intentionally  added  MNPs  to 
products, providing new evidence for better preventive 
policies. CUSP comprises five new projects running from 
2021 to 2025:

POLYRISK  examines  the  toxic  effects  of  micro-  and 
nanoplastic  particles  (MNP)  on  the  immune  system. 
Using  advanced  methods  to  chemically  detect  and 
quantify MNP, the projects aims to understand the key 
mechanisms of MNP toxicity in vitro, and find biomarkers 
of toxicity in blood and saliva.

AURORA (Actionable eUropean ROadmap for early-life 
health  Risk  Assessment  of  micro-  and  nanoplastics) 
aims  to  develop  a  framework  for  risk  assessment  to 
evaluate  the  impact  of  microplastic  and  nanoplastic 
pollution during pregnancy and early life, using in-depth 
testing and epidemiological data to reveal impacts on 
child development and health.

IMPTOX will develop an innovative analytical platform 
to  investigate  the  effect  and  toxicity  of  micro- 
and  nanoplastics  combined  with  environmental 
contaminants on the risk of allergic disease in preclinical 
and clinical studies.

PLASTICHEAL  aims  to  develop  new  methodologies 
and  reliable  scientific  evidence  for  regulators  to  set 
the knowledge basis for adequate risk assessment of 
micro- and nanoplastics. It will investigate both short- 
and long-term potential health impacts. 

PlasticsFatE  (Plastics  Fate  and  Effects  in  the  human 
body)  will  implement  a  comprehensive  measurement 
and  testing  programme  to  improve  and  validate 
methods and tools for the identification of micro- and 
nanoplastics.  It  will  lead  to  a  new  risk-assessment 
strategy  supporting  the  health-relevant  aims  of  the 
European strategies for plastics.

Box 3: EU projects

A  number  of  European  projects  are  underway  that 
will  advance  understanding  of  nanoplastic  behaviour, 
hazards and potential remediation, including:

CE4Plastics4 
(2022-2024)  -  advancing  methods 
for  identification  and  quantification  of  nanoplastics 
released to drinking water from single-use and reusable 
plastic bottles, and raising awareness in the population 
about nanoplastic risks.

VORTEX5 
(2018-2023) 
transformation of plastic in the ocean.

- 

investigating  microbial 

In-No-Plastic6  (2020-2023)  -  developing  technology 
and  social  initiatives  to  remove  plastic  from  aquatic 
ecosystems.

MIGMIPS7  (2021-2023)  -  using  gut  barrier  models 
to  study  the  translocation  of  microplastics  and 
nanoplastics.

MS4Plastics8 (2022-2024) - developing innovative pre-
treatment  and  pre-concentration  protocols  to  detect 
and  characterise  low  µm-range  microplastics  and 
nanoplastics (1 nm – 20 µm) in edible fish and shellfish 
samples.

The  EUROqCHARM9  consortium  is  evaluating  existing 
methodologies  for  plastic  pollution  assessment,  with 
the aim of harmonising them on a European level. The 
training  network  MonPlas10  (Monitoring  micro-  and 
nanoplastics),  meanwhile,  is  working  with  early  stage 
researchers to develop technology for detecting plastic 
particles in water, in order to provide standardised data. 

These  initiatives  complement  the  EU’s  RESPONSE11 
project (2020–2023): Towards a risk-based assessment 
in  marine  ecosystems. 
of  microplastic  pollution 
RESPONSE  integrates  expertise  on  oceanography, 
environmental  chemistry,  ecotoxicology,  experimental 
ecology  and  modelling  to  answer  key  research 
questions  on  fate  and  impact  of  microplastics  (MPs) 
and nanoplastics (NPs), aiming to provide ecologically 
relevant  strategies  for  assessing  the  distribution 
pathway  and  biological  effects  of  plastic  particles  in 
marine ecosystems. 

https://cordis.europa.eu/project/id/101059423
4 
https://cordis.europa.eu/project/id/772923
5 
https://cordis.europa.eu/project/id/101000612
6 
https://cordis.europa.eu/project/id/101032657
7 
https://cordis.europa.eu/project/id/101023205
8 
9 
https://www.euroqcharm.eu/en
10  https://www.monplas.eu/project
11  https://www.response-jpioceans.eu/
12  https://cusp-research.eu/

16

1.2 Regulatory landscape 

The  European  Chemicals  Agency  (ECHA)  estimate 
that, each year, around 42 000 tonnes of intentionally 
added microplastics end up in the environment. The 
largest single source of microplastic pollution is the 
granular infill material used on artificial turf pitches, 
with releases of up to 16 000 tonnes. On the 30th 
August  2022  the  European  Commission  released  a 
draft proposal regarding the restriction of synthetic 
polymer microparticles13. 

The  restriction  in  the  draft  proposal  would  comprise 
synthetic  polymer  microparticles  below  5  mm,  and 
fiber-like particles below 15 mm, that are intentionally 
used  in  products,  and  may  result  in  environmental 
release.  It  could  ban  the  use  of  microplastics  in 
cosmetics,  cleaning  products,  pesticides,  and  sports 
fields  amongst  others,  and  thus  covers  emissions  of 
microplastics  that  are  preventable.  According  to  the 
draft,  the  “Commission  considers  it  appropriate  to 
exclude natural, degradable, and soluble polymers from 
the definition of synthetic polymer microparticles.” 

Box 4: Measures in the proposed restriction on microplastics in products 
placed on the EU/EEA market

On  August  30th  2022  the  European  Commission 
released  a  draft  proposal  for  a  restriction  on 
microplastics.  The proposal does not include a lower 
limit  on  the  size  of  microplastics,  to  incorporate 
all  smaller  sized  plastic  particles  –  thus  including 
nanoplastics. In addition, the proposal also adjusted 
the  timings  for  companies  to  adapt  to  the  law  – 
allowing  some  cosmetic  products  up  to  12  years 
to  comply.  It  is  worth  noting  that  at  present,  it  is 
also stated in this draft that enforcement cannot be 
undertaken on plastic particles below 1µm, due to 
a lack of suitable detection techniques for particles 
this small .

The  proposed  ECHA  restriction,  which  has  not  yet 
been adopted by the European Commission, includes 
several measures detailed below: 

 y A restriction on the placing on the market 
of  microplastics  (and  nanoplastics)  on 
their own, or in mixtures, where their use will 
inevitably result in releases to the environment, 
irrespective of the conditions of use. Examples 
are  cosmetics,  cleaning  and  laundry  products, 
fertilisers,  plant  protection  products  and  seed 
coatings. For some of these uses, a transitional 
period  is  proposed  to  allow  sufficient  time  for 

stakeholders to comply with the restriction. For 
example,  a  ban  on  artificial  turf  microplastic 
infill, after six years, is proposed for this material.

 y A 

labelling  requirement  to  minimise 
releases  to  the  environment  for  uses  of 
microplastics (and nanoplastics) where they 
are not inevitably released to the environment 
but where residual releases could occur if they 
are  not  used  or  disposed  of  appropriately  (e.g. 
paints  and  inks).  These  uses  are  not  proposed 
to be prohibited but would need to be reported 
to  ECHA  to  ensure  that  residual  releases  are 
monitored and could be controlled in the future, 
and  suppliers  would  also  be  obligated  to  give 
instructions  on  how  residual  releases  can 
be  minimised.  This  provision  will  also  enable 
information exchange along the supply chain.

 y A  reporting  requirement  to  improve  the 
quality of information available to assess 
the potential for risks for remaining uses 
of microplastics (and nanoplastics) in the 
future. This is considered a cost-effective way 
to enable the Commission and Member States 
to  consider  if,  and  to  what  extent,  additional 
action could be needed in 5-10 years.

Source: ECHA, 2019;  European Commission draft proposal regarding the restriction of synthetic polymer 
micropartcles, Comitology register (europa.eu)14 

13  https://ec.europa.eu/transparency/comitology-register/screen/documents/083921/1/consult?lang=en
14  https://ec.europa.eu/transparency/comitology-register/screen/documents/083921/1/consult?lang=en

 
The adoption of any lower particle size limit in the 
scope of restriction, and the level at which this was 
set,  would  clearly  have  large  implications  for  the 
regulation  of  nanoplastics.ECHA’s  initial  opinion 
included  such  limit  of  100nm,  which  however 
ECHA’s  Committee  for  Risk  Assessment  (RAC) 
did  not  support  (RAC/SEAC,  2020a;  RAC/SEAC, 
2020b) while in contrast, the Committee for Socio-
economic Analysis (SEAC) included a lower size limit 
of 1 nm and recommended a temporary lower size 
limit of 100 nm to ensure that the restriction can 
be enforced by detecting microplastics in products 
(ECHA,  2020a).  The  2022  draft  Commission 
proposal  does  not  include  a  lower  limit  on 
microplastic  size,  thus  incorporating  nanoplastics 
within  this  restriction.  However,  it  includes  the 
text  stating  that  where  the  concentrations  of 
microparticles  cannot  be  determined  by  current 
analytical  methods  (available  at  the  time  of 
enforcement)  only  particles  of  0.1µm  or  above 
shall be considered – for microfibres this is 0.3 µm 
in length (EC, 2022b).

Specific  transitional  periods  are  outlined  in  the 
draft  proposal  released  in  August  2022  by  the 

1.2.2 Other plastics relevant regulation

The  ECHA’s  restriction  dossier15  identifies  the 
role  of  plastic  additives  (such  as  fillers,  UV 
stabilisers  and  plasticisers)  in  the  (eco)toxicity  of 
microplastic  (and  nanoplastics)  as  an  important 
data gap. The authors also suggest that the current 
risk  assessment  of  these  substances  is  unlikely 
to  have  considered  exposure  of  organisms  via  a 
microplastic (or nanoplastic) vector (ECHA, 2019).

A  mapping  exercise  by  ECHA  and  the  chemicals 
industry  identified  a  list  of  over  400  functional 
additives or pigments used in plastics, as well as 
the  polymers  in  which  they  are  most  commonly 
found,  and  typical  concentration  ranges  (ECHA, 
n.d.).  Focusing  on  plasticisers,  flame  retardants, 
pigments, 
agents, 
nucleating  agents  and  types  of  stabilisers,  the 
mapping included substances registered by REACH 
at  production  above  100  tonnes  per  year.  This 
figure is a fraction of the total number of plastic 

antioxidants, 

antistatic 

17

transitional 

EC,  to  allow  stakeholders  time  to  find  suitable 
alternatives.  These 
time  periods 
included 3-12 year compliance periods, depending 
on  the  context  of  the  diverse  products  listed  – 
which  ranged  from  agricultural  plant  protection 
pesticides,  to  cosmetic  products  such  as  makeup 
(EC, 2022b).

in 

formed  microplastics 

Other  options  for  reducing  the  releases  of 
unintentionally 
the 
aquatic  environment  are  being  considered  by  the 
Commission  as  part  of  its  Plastics  Strategy  and 
the new Circular Economy Action Plan, concerning 
“microplastics  created  during  the  lifecycle  of  a 
product through wear and tear or emitted through 
accidental spills” and provisions to reduce the use 
of single-use plastic products have been adopted 
as  part  of  the  Single  Use  Plastic  Directive  (EU 
2019/904) .

All  these  initiatives  include  nanoplastics  as 
part  of  microplastics.  They  do  not  include 
any  specific  measures  on  nanoplastics, 
where knowledge to support such legislation 
remains lacking. 

additives  that  have  been  used  –  a  more  recent 
review expanded on the ECHA mapping and found 
more  than  6  000  chemicals  reported  in  plastics 
(Aurisano, Weber and Fantke, 2021). Of these, over 
1  500  were  identified  as  ‘of  concern’,  meaning 
they should be prioritized for substitution by safer 
alternatives.

restrictions  placed  on  some 

Since  the  implementation  of  REACH  in  2007, 
more  than  50  plasticisers  have  been  registered 
(European 
and 
Plasticisers,  n.d.).  For  example,  from  February 
2015,  DEHP,  BBP,  DBP  and  DIBP  plasticisers  were 
no longer allowed to be produced in the EU, unless 
authorisation had been granted for a specific use. 
Some  restrictions  precede  REACH;  e.g.  phthalates 
(which  can  be  used  as  plasticisers)  in  children's 
toys  have  been  restricted  in  the  EU  since  1999. 
(Section  4.7  gives  more  detail  on  plastic  additive 
toxicity.) The Extended Producer Responsibility 

15  Registry of restriction intentions until outcome - ECHA (europa.eu): https://echa.europa.eu/registry-of-restriction-intentions/-/

dislist/details/0b0236e18244cd73

18

(EPR) principle (as contained in Council Directive 
2008/98/EC16  on  waste  management)  aims  to 
transfer  the  elimination  costs  of  used  products 
from the taxpayer to the consumer, by integrating 
them in the sales prices of the new products. Each 
Member  State  may  introduce  the  EPR  concept 
into  its  own  legal  framework,  and  decide  how  to 
encourage  manufacturers  to  participate  in  the 
prevention, re-use, recycling and recovery of used 
plastic products. As one example of implementation, 
national collecting schemes (NCS) for agricultural 
waste have been implemented in several Member 
States,  with  several  countries  encouraging  users, 
distributors and producers to recycle agri-plastics 
(Ireland,  Iceland,  Sweden,  France,  Spain,  Norway, 
Germany and also the UK).

The  European  Commission  is  also  planning  to 
develop  a  proposal  for  the  registration  of 
selected polymers in Europe by 2022. At present, 
neither  registration  nor  evaluation  of  polymers 
has  been  required  by  REACH,  but  pressure  is 
mounting  worldwide  to  limit  plastic  entry  into 
the  environment,  and  registration  of  polymers  is 
already  required  in  the  USA,  Japan,  South  Korea 
and  China.  Citing  the  generation  of  micro-  and 
nanoplastics  as  an  “inherent  hazardous  property 
of  plastic  polymers”  amongst  other  risks,  some 
NGOs have called for all synthetic polymers to be 
included in the pre-registration submissions under 

REACH  (ChemSec,  2020).  They  also  note  that 
plastics  excluded  from  the  proposed  EU  safety 
checks include polystyrene, which has been linked 
to lung inflammation in rats, and polyacrylamides, 
which can degrade to the monomer acrylamide, a 
neurotoxin (EEB, 2021). 

The  updated  Drinking  Water  Directive17,  which 
entered  into  force  in  January  2021,  addresses 
microplastics and endocrine disruptors, introducing 
minimum  hygienic  requirements  for  materials  in 
contact with water intended for human consumption 
in the EU. Any material that is intended to be used 
in new installations or, in the case of repair works 
or reconstruction, in existing installations and that 
comes  in  contact  with  drinking  water  between 
the abstraction points and the tap will need to be 
conform with the hygienic requirements.

It  is  important  to  recognise  that  plastic  pollution 
and  the  risk  from  nanoplastics 
is  a  global 
problem.  At  the  UN  Environment  Assembly  in 
2022,  173  countries  agreed  to  develop  a  legally 
binding  treaty  on  plastics.  Legislation  in  North 
America is beginning to be enacted: California has 
spearheaded  measures  to  prevent  microplastics 
entering the environment, and Canada is banning 
single use plastics, however there are currently no 
regulations specifically addressing nanoplastics.

16  http://eur-lex.europa.eu/LexUriServ/LexUriServ.do?uri=CELEX:32008L0098:EN:NOT
17  https://ec.europa.eu/environment/water/water-drink/review_en.html

19

2. Detection and assessment 

Nanoplastics have now been detected at the North 
and  South  Poles,  in  remote  lakes  in  the  Northern 
Hemisphere, and in snow in the Austrian Alps (e.g. 
Materić  et  al.,  2020;  discussed  further  in  section 
3.4).  They  can  be  traced  within  biological  cells, 
and  measured  in  microscopic  organisms  (see 
3.5  and  5.2).  However,  techniques  for  detecting 
nanoplastics 
in  environmental  and  biological 
samples  are  in  their  infancy,  along  with  ways  of 
assessing  their  fate  and  effects  on  ecosystems, 
biota and humans.

Suitable methods must be developed, standardised 
and validated to meet the challenge of assessing 
the types, amounts, and sizes of fragmented plastic 
particles  in  environmental  samples  –  something 
that remains limited in even comparatively ‘clean’ 
samples  such  as  domestic  water  (Schwaferts  et 

al.,  2019;  Valsesia  et  al.,  2021).  When  samples 
are  more  complex,  for  example  containing  large 
amounts of biological material, matrix effects and 
contamination become a greater issue, making an 
already challenging task even more so. 

Isolating  nano-sized  plastics  from  the  huge 
numbers  of  natural  particles  of  similar  sizes 
found  in  ecosystems  is  one  challenge;  assessing 
their  persistence  –  of  utmost  importance  for 
evaluation of their risk potential – is another. This is 
pertinent  to  assessing  biodegradability.  European 
standards  exist  for  assessing  the  compostability 
of plastics in industrial composting plants and the 
biodegradability  of  mulch  films  in  soil  for  use  in 
agriculture, for instance, but there is no standard for 
plastic degradability in water, or within organisms, 
as yet (EEA, 2020; EC, 2020).

Garbage in the oceans kills live animals (©Shutterstock, photo by HilaryDesign)

20

2.1 Methods of characterisation and quantification

them 

Detecting  and  quantifying  nanoplastics  in  the 
environment requires methods that can distinguish 
from 
nanoparticles  and  separate 
the 
background  matrix.  Methods  of 
isolating  and 
characterising  nanoplastics  utilising  dynamic  light 
scattering  and  Raman  microscopy,  for  example, 
are  under  development,  but  adapting  commonly 
used  techniques  used  to  study  microplastics  can 
be problematic; conventional filtration, for example, 
would be excessively time-consuming and costly, as 
“thousands to tens of thousands of litres of water 
would  have  to  be  filtered  through  nanoscale-sized 
pores to acquire statistically relevant quantities of 
environmental aquatic nanoplastics” (Merzel et al., 
2020). Schwaferts et al. (2019) provided an overview 
of  methods  that  are  applied  to  microplastics,  that 
may also be used for nanoplastics (Table 2). 

Being  composed  mainly  of  carbon,  meanwhile  – 
not too different from organic matter – means that 
nanoplastics  evade  many  available  instrumental 
detection techniques (Jiménez-Lamana et al., 2020; 
Lehner et al., 2019). Additionally, it is unknown what 
plastic  debris  will  look  or  behave  like  throughout 
the 
further 
uncertainty  into  what  to  search  for  (EC,  2017).  

degradation 

introducing 

cycle, 

Many  techniques  are  able  to  detect  and  size 
nanoparticles, but detecting nanoplastics according 
to  size  is  insufficient  –  chemical  characteristics 
must be identified to classify particles as plastics. 
Meanwhile,  a  technique  for  detecting  one  type 
of  nanoplastic  may  not  work  for  a  different  type, 
without adjustment (Correia and Loeschner, 2018). 

(DLS), 

light  scattering 

Some  of  the  methods  that  show  promise  include 
dynamic 
  microscopy, 
nanoparticle tracking analysis, field flow fractionation 
(FFF) techniques and mass spectrometry. Techniques 
are also being developed using doped nanoplastics, 
whereby particles enriched with detectable metals 
(e.g. lead, palladium) are added to complex matrices 
and  searched  for  using  robust  methods  based  on 
mass  spectrometry  (MS)  (Mitrano  et  al.,  2019). 
Cassano  et  al.  (2021)  proved  able  to  produce 
stable,  high-throughput,  traceable  polypropylene 
nanoplastics by introducing selected markers such 
as  metallic  nanoparticles,  fluorescent  organic  and 
metal-organic  dyes  or  quantum  dots  –  relevant 
for  toxicity  and  biodistribution  studies.  At  a  larger 
scale, Facchetti et al. (2020) successfully identified 
and characterised fluorescent and metal-doped PVC 
microplastics 
(PVC-Platinum  octaethylporphyrin 
(PtOEP) microplastics in the size range of 100 µm) 
in mussels using fluorescence microscopy. 

Confocal Raman imaging microscope, ©Wikipedia Commons CC BY 4.0

 
21

Table 2: Methods for characterising particle size and morphology Adapted from Schwaferts et al. (2019).

Particle characterisation by light scattering

Applying the scattering of laser light on particles to obtain information on physical properties. 
Generally easy application, and able to gather information including size, particle size distribution, 
surface charge, stability, aggregation behaviour, and concentration.

Methodology

Range/Limits

Advantages

Disadvantages

Dynamic Light 
Scattering – DLS

1 nm–3 µm conc.  
10-6 –10-1

Fast, cheap, in 
situ, non-invasive, 
aggregation, direct 
coupling

Multi Angle Light 
Scattering – MALS

10 nm–1000 nm

Online coupling

Large particles 

Polydispersity

Complex matrix

Non-spherical 
particles

Requires clean 
samples

Laser Diffraction – LD

10 nm–10 mm  
conc. 10-5 –10-1

Large size range, 
easy, fast, automated

Only spherical model

Nanoparticle Tracking 
Analysis – NTA

30 nm–2 µm 
conc. 10-6 –10-5

Particle imaging techniques 

Better with 
polydisperse samples, 
complex media, 
particle corona

Complex in operation

Spherical model

Microscopy is the most viable method for obtaining information on a sample’s morphology. Three 
groups of microscope operation modes are most prominent in particle imaging: optical, electron 
(EM) and scanning probe (SPM).

Methodology

Range/Limits

Advantages

Disadvantages

Transmission Electron 
Microscopy – TEM

<1 nm

High resolution, 
precise size 
information

Quantification difficult

Sample preparation

Expensive

Quantification difficult

Scanning Electron 
Microscopy – SEM

ca. 3 nm

High resolution

Sample preparation

Charging effects

22

Environmental 
Scanning Electron 
Microscopy – ESEM

ca. 30 nm

Wet samples, 
environmental 
conditions, non-
conductive samples

Energy Dispersive 
Spectroscopy – EDS

nm range

Complementary to EM

Atomic Force 
Microscopy – AFM

ca. 0.1 nm

High resolution, AFM-
IR, TERS, In liquid, 
high resolution

Reduced resolution

Elemental information 
not sufficient

Slow

Small area

Artefacts due to 
particle movement

Conductive samples

Scanning Tunnelling 
Microscopy – STM

Confocal Laser 
Scanning Microscopy 
– CLSM

Near-field Scanning 
Optical Microscopy – 
NSOM

ca. 1 nm

High resolution

Slow

>0.2 µm

Fluorescence imaging

ca. 30 nm

Fluorescence

Small area

Small area

Diffraction limit

Slow

Small area

Chemical identification and characterisation

Chemical identification provides information on type of particle, additives, and ageing. This has 
usually been achieved by combining vibrational spectroscopy with optical microscopy for imaging. It 
becomes increasingly difficult with decreasing particle size. 

Methodology

Range/Limits

Advantages

Disadvantages

Atomic Force 
Microscopy Infrared 
Spectroscopy – AFM-
IRIR

>50 nm

High resolution, 
chemical imaging

Slow, small area

Electrophoretic Light 
Scattering – ELS

1 nm–3 µm

Fast, cheap, non-
invasive, and may be 
integrated with DLS 
instrument.

Particle characteristic 
is sensitive to 
environment (pH)

23

Raman Microscopy – 
RM

>0.5 µm

Non-destructive, easy 
sample preparation, 
fast, no interference 
from water.

Fluorescence

LOD dependent on 
polymer type

Some polymers 
difficult

Dry sample needed

Preconcentration 
necessary

Early-stage research 
based on well-
characterised samples 
and controlled plastics 
– mammalian tissue 
samples were spiked 
with nanoplastic beads. 
Cannot distinguish 
polymer particles from 
other particles (ACS, 
2020).

Pyrolysis Gas 
Chromatography 
Mass Spectrometry – 
Py-CG-MS

Thermal Extraction 
Desorption Gas 
Chromatography 
Mass Spectrometry – 
TED-GC-MS

Limits of detection 
(LOD): 0. 1µg – 1µg

Little sample 
preparation.

Measurement with 
matrix, higher sample 
masses.

Other methods of separation and combined techniques

Flow cytometry (FC)

Down to 200 nm 
(spatial detectable 
limit) in freshwater

Field Flow Fractionation 
(a family of methods, 
including Asymmetrical 
Field Flow Fractionation 
(AF4), Centrifugal Field 
Flow Fractionation 
(CF3) and Thermal FFF

1nm (AF4); ca. 50nm 
(CF3);

High potential for 
the processing of 
environmental samples

Particle loss and 
sample alteration 
or aggregation are 
minimized 

Thermal FFF can 
separate different 
polymers of the same 
molecular weight 
(Postnova, n.d.)

24

Other MS-based 
techniques, e.g. liquid 
chromatography 
coupled to high-
resolution mass 
spectrometry LC-
HRMS, thermal 
fragmentation and 
matrix-assisted laser 
desorption/ionisation 
time-of-flight mass 
spectrometry MALDI-
TOF MS, and Single 
particle inductively 
coupled plasma mass 
spectrometry SP-ICP-
MS.

LC-HRMS was recently 
proposed for the 
mass quantification of 
nanoplastics and small 
microplastics

SP-ICP-MS has been 
developed based on 
model studies and 
synthetic samples of 
polystyrene plastic 
debris; “the study of 
more complex matrices 
and real environmental 
samples [would require] 
further analytical 
development.”

(Jiménez-Lamana et 
al., 2020)

Fig. 4. Centrifugal field flow fractionation allows particles of the least mass to be separated from heavier, larger 
particles. Source: Postnova.com 

Box 5: Examples of methodologies used in detecting and identifying 
nanoplastics 

Fourier transform infrared (FTIR) and Raman 
spectroscopy:  These  methodologies  are  widely 
applied to microplastic detection (e.g. Zhang et al., 
2019)  and  show  some  promise  at  the  nanoscale, 
particularly Raman spectroscopy. FTIR obtains an 
infrared spectrum of absorbed or scattered light/

energy of a sample. Molecules can be identified by 
their  characteristic  absorption  spectrum.  Raman 
spectroscopy also uses scattered light to measure 
vibrations  of  molecular  bonds.  Both  techniques 
measure a different type of molecular fingerprint.

25

Surface-enhanced  Raman  spectroscopy  has  been 
used  to  detect  nanoplastics.  Lv  et  al.  (2020) 
applied this technique to pure water and seawater 
samples,  and  successfully  detected  nanoplastics 
of  100  nm  in  size  down  to  concentrations  of  40 
µg/mL.  Valsesia  et  al.  (2021)  also  used  Raman 
spectroscopy  to  analyse  nanoplastics 
in  the 
tissues  of  salt-water  mussels,  a  technique  that 
drove  nanoplastics  to  self-assemble  and  self-
isolate.  Their  process  involved  eliminating  the 
biological  matrix  with  enzymatic  digestion  and 
filtering, followed by collection of nanoplastics on 
a  surface  micro-machined  with  sub-micron-sized 
cavities.  As  the  solution  containing  nanoparticles 
dried,  capillary  action  assembled  the  suspended 
particles  into  the  cavities,  isolating  and  trapping 
them across the surface, enabling analysis. 

Ultrafiltration  with  dynamic  light  scattering 
(DLS):  DLS  measures  fluctuations  in  scattered 
light  to  determine  the  diffusion  coefficient  –  and 
subsequently  size  –  of  particles.  Ter  Halle  et  al. 
(2017) used ultrafiltration combined with dynamic 
light  scattering  to  detect  nanoparticles  in  water 
samples from the North Atlantic Subtropical Gyre, 
and  subsequently  ascertained  their  chemical 
composition  (plastics)  via  pyrolysis  coupled  with 
gas chromatography-mass spectrometry. 

(GC-MS): 

Pyrolysis coupled with gas chromatography-
In  pyrolysis, 
mass  spectrometry 
organic material is heated to high temperatures in 
the absence of oxygen (to avoid combustion risk) 
until it decomposes. These decomposition products 
are  then  separated  via  gas  chromatography  (GC) 
and  characterised  by  mass  spectrometry  (MS)  at 
a  molecular  level,  to  reveal  the  quantities  and 
compositions  of  the  substances  present.  This 
technique has been used to identify nanoparticles 
as nanoplastics (Ter Halle et al., 2017).

Hyperspectral  imaging:  This  technique  collects 
many  images  at  different  wavelengths  for  the 
same  spatial  area,  that  can  be  analysed  for 
individual  materials.  Mattsson  et  al.  (2017)  used 
hyperspectral  imaging  to  detect  and  characterise 
polystyrene  nanoparticles  with  a  diameter  of  52 
nm  in  the  brain  tissue  of  freshwater  carp,  having 
been  passed  up  the  food  chain  from  the  small 
planktonic crustacean Daphnia magna. 

techniques 

Acquisition 
imaging, 
for 
visualized  as  sections  of  the  hyperspectral  datacube  with 
its two spatial dimensions (x,y) and one spectral dimension 
(lambda) - ©Wikipedia Commons CC BY 4.0

hyperspectral 

Field  Flow  Fractionation:  Some  researchers 
have  highlighted  the  potential  of  field  flow 
fractionation  methods  (FFF4)  for  separating  out 
nano-sized  particles  in  a  sample  (Gigault  et  al., 
2017;  Schwaferts  et  al.,  2020;  Loeschner  et  al., 
2022).  Asymmetrical  FFF  (AF4)  and  Centrifugal 
FFF (CF3; Fig. 4) use a physical force to separate 
particles  as  they  flow  along  a  channel,  based 
on  their  size  or  mass.  Particles  can  be  further 
characterised  by  coupling  these  techniques  with 
other methods such as Raman spectroscopy and 
multi-angle laser light scattering. In Thermal FFF, 
a  temperature  gradient  is  the  driving  force  of 
separation, with particles are separated according 
to their mass and chemical composition.

Loeschner et al. (2022) note that sample preparation 
needs  careful  consideration  when  looking  for 
nanoplastics  with  FFF.  Suspensions  should  not 
use  strong  acids  or  bases,  for  example,  which 
could  affect  nanoplastics.  Very  few  studies  have 
attempted to apply FFF for analysing nanoplastics 
in a complex matrix. Wahl et al. (2021) used AF4 in 
conjunction with pyrGC-MS to detect nanoplastics 
in  soil;  Correia  and  Loeschner  (2018)  applied 
AF4  and  Multi-angle  laser  light  scattering  to  fish 
samples spiked with PS and PE nanoplastics.

26

Box 6: Core challenges

Fig. 5. Challenges and pitfalls in micro- and nanoplastics research. Source: Paul et al. (2020) 

Some  of  the  challanges  to  be  addressed  in 
developing  methodologies  for  detecting  and 
anlaysing nanoplastics include:

 y Complex  matrices:  samples 

range 

from 
‘clean’  such  as  drinking  water, 
(river/sea/lake), 
alimentary,  

sediment, 

soil, 

environmental  water 

relatively 
to 
wastewater, 
and biota.

 y Many 

different  materials 

identify: 
numerous plastics are used in global markets. 

to 

2.2 Considerations for policy

to 

is  possible 

It 
restrict  manufacturing  of 
nanoplastics,  addressing  the  release  of  primary 
nanoplastics  (as  in  the  restriction  on  intentionally 
added microplastics discussed in 1.2.1, for example). 
However,  secondary  nanoplastics  can  only  really 
be  addressed  through  policies  aimed  at  macro 
and  microplastics  –  as  in  the  European  Strategy 
for  Plastics  in  a  Circular  Economy,  for  example, 
or  through  the  European  Commission’s  initiative 

 y No  reference  material,  and  no  clearly  defined 

legal definition or standard metrics.

 y Fluorescence/label  leakage:  markers  used  on 
nanoplastics may confound findings (see 3.5.2).

 y Pristine  vs.  aged  articles:  heterogeneity  in 
target particles increases analytical complexity.

 y Avoiding 

contamination 
extraction, concentration and handling.

during 

sample 

 y Improving resolution, sensitivity and specificity 

in analysis.

on  reducing  un-intentional  microplastic  pollution 
into  the  environment.  Increasing  recycling  rates, 
regulating  single-use  conventional  plastics  (e.g.  as 
in  the  Plastic  Bags  Directive18),  finding  safe  ways 
of dealing with end-of-life plastics, and developing 
biodegradable plastics that pose less risk to health 
and  ecosystems  than  conventional  plastics,  are 
overarching  goals  that  may  address  the  ongoing 
In  regulation, 
secondary  nanoplastic  pollution. 

18  https://environment.ec.europa.eu/topics/plastics/plastic-bags_en

27

it  may  also  be  useful  to  target  certain  types  of 
plastic that are more likely to produce or persist as 
nanoplastics – or be more hazardous, either due to 
their  inherent  characteristics  and  toxic  effects,  or 
their  chemical  additives.  Remediating  secondary 
nanoplastics already in the environment is another 
challenge altogether.

Meanwhile,  the  difficulty  in  detecting  nanoplastics 
presents problems for monitoring the effects of any 
restrictions on nanoplastic levels in the environment. 
Currently  available  methodologies  require  greater 
sensitivity, greater ease of use, wider applicability, and 
often must be combined in order to fully characterise 
the nanoplastic content of a complex sample.

Additionally,  methods  must  be  developed  that 
‘mark’  plastics  and  causally  link  these  to  origin, 
source  or  manufacturer;  and  models  are  needed 
to  assess  “the  relationships  between  polymer 
structural  characteristics  and  the  formation  of 
smaller  plastic  nanoparticles  in  nature,  due  to 
embrittlement,  fragmentation  or  degradation” 
(SAPEA,  2019).  Overall,  harmonised  protocols  for 
monitoring the concentrations and compositions of 
marine plastics must be developed and overcome 
three main issues, say Setälä et al. (2019): “how to 
carry out the field sampling, how to eliminate other 
particulate matter in the sample without harming 
the plastic itself, and how to accurately identify the 
particles”, all without introducing contamination to 
the sample in the process.

To advance the detection, analysis and monitoring of 
nanoplastic pollution, Horizon 2020 is supporting the 
European  Cluster  on  Health  Impacts  of  Micro-  and 
Nanoplastics  (2021–25),  a  group  of  five  research 
initiatives (see Box 3, section 1.1.3). These initiatives 
complement 
the  EU’s  RESPONSE  project 
(2020–2023):  Towards  a  risk-based  assessment 
of  microplastic  pollution  in  marine  ecosystems. 
RESPONSE  integrates  expertise  on  oceanography, 
environmental chemistry, ecotoxicology, experimental 
ecology  and  modelling  to  answer  key  research 
questions  on  fate  and  impact  of  microplastics 
(MPs)  and  nanoplastics  (NPs)”,  aiming  to  provide 
“ecologically  relevant  strategies  for  assessing  the 
distribution pathway and biological effects of plastic 
particles in marine ecosystems”. 

A  key  consideration  of  future  policy  is  safety,  with 
many  core  research  questions  aiming  to  address 
the level and type of risk that nanoplastics pose to 
global  ecosystems,  the  marine  environment,  and, 
crucially,  human  health.  While  nanomaterials  and 
primary nanoplastics can be evaluated to assess if 
they are “safe by design” across their lifecycle (Allan, 
Sokull-Kluettgen and Patri, 2020), this is less feasible 
for  secondary  nanoplastics  –  which  represent  the 
majority of nano-sized plastic particles found in the 
environment. As attention has turned to nanoplastics, 
numerous  stakeholders,  including  plastic  producers, 
have  developed  strategies 
to  combat  plastic 
pollution,  but  much  work  remains  to  address  data 
gaps pertaining to nanoplastics.

Box 7: Methodological challenges and policy

Robust,  accurate,  harmonized  methodology  for 
detecting  and  analysing  nanoplastics  in  different 
media (e.g. biological or environmental samples) is 
needed to enable various aspects of policy relating 
to nanoplastics:

 y Evaluate  effects  and  safety  of  different 

nanoplastics, including biodegradable

 y Assess  hazard  posed  by  nanoplastics  in  the 

environment

 y Direct 

restrictions  at  most  hazardous 

 y Establish levels in different environments

nanoplastics

 y Monitor changes to levels in the environment

 y Estimate effects of regulations on occurrence 

 y Trace origin of nanoplastics in the environment

of nanoplastics

 
28

3. Nanoplastics in the environment

Primary  and  secondary  nanoplastics  may  arise 
in  terrestrial,  and  aquatic  environments,  via  a 
number  of  key  sources  –  including  agriculture 
and  fishing.  Wastewater  treatment  does  not 
remove  all  nanoplastic  particles.  Once  they  enter 
the  environment, 
transport, 
and  fate,  depends  on  a  number  of  factors  –  for 
example, plastics denser than water will sink, while 
others  float.  Nanoplastic  particles  are  subject 
to  fragmentation  and  degradation  processes, 
and  there  are  interesting  new  findings  about 
biotransformation  that  occurs  during  digestion  in 
some organisms. 

their  behaviour, 

Ingestion  of  microplastics  by  aquatic  organisms 
in  many  studies,  and  the 
has  been  shown 
literature  on  nanoplastic  exposure  via  this  route 
is  growing,  together  with  evidence  that  particles 
may  translocate  (move  to  different  organs  and 
tissues) within an organism. In terrestrial habitats, 
nanoplastics  may  move  downwards  through  soil 
and  be  taken  up  into  plant  and  fungus  tissues, 
with potential implications for trophic transfer and 
ecosystem functioning.

3.1 Sources and routes into the environment

Most  plastics  are  produced  and  used  on  land, 
and  hence  the  sources  for  microplastic  and 
nanoplastics  are  mainly  land-based.  However,  as 
these  small  plastic  particles  are  uniquely  light, 
persistent, and able to remain dispersed in water 
– they can travel far across the planet. Over 800 
million  tonnes  of  plastic  pollution  in  the  sea  is 
estimated  to  come  from  land-based  sources,  as 
well  as  fishing,  other  aquaculture  activities,  and 
coastal  tourism  (Yee  et  al.,  2021).  There  are  also 
some sea-based sources: ‘ghost gear’ – lost fishing 

equipment  (nets,  ropes,  lines)  makes  up  10%  of 
marine litter – with between 500 000 and 1 million 
tonnes  left  in  the  ocean  every  year  (WWF,  2020). 
This  plastic  pollution  makes  up  half  of  the  ’Great 
Pacific Garbage Patch’ which is floating in the North 
Pacific Ocean (and is already deadly to marine life 
in  its  macro  form,  through  entanglement).  When 
degraded  to  micro-  and  nanoplastic  size,  these 
plastics  can  also  be  ingested  by  marine  fauna  – 
possibly entering the food chain. 

29

Fig. 6. Sources and fate of micro- and nanoplastics in the environment. Both micro- and nanoplastics can occur 
in  both  aquatic  and  terrestrial  environment  and  undergo  a  range  of  abiotic  and  biologicalbiologicalbiosolid 
removalbiological degradation processes, and eventually enter the food chain, water supplies, and the air. There 
is at present only a speculative link between the environment and ingestion and uptake of these plastic particles 
in the human body. Source: Yee et al. (2021)

30

Wastewater treatment processes may remove more 
than 95% of microplastics, but there are currently 
few  research  findings  on  levels  of  nanoplastics 
remaining in treated water; most studies have only 
quantified  particles  greater  than  1  µm  (Ali  et  al., 
2021). One experiment, using nanoplastics marked 
with palladium (as well as microplastic particles), 
found  that  plastic  removal  is  strongly  correlated 
with suspended solids removal – at least 98% of 
particulate plastics were associated with solids in 
the  study  (Frehland  et  al.,  2020).  The  researcher 
acknowledged  that  it  is  not  currently  feasible  to 
measure  nanoplastics  in  a  full-scale  wastewater 

treatment plant, but their findings using a simplified 
model  suggest  that  in  real-life  treatment  –  with 
additional processing steps – even higher levels of 
solids and plastic particles can be retained.

Any particles that do remain in treated water will 
be  introduced  into  rivers  and  oceans,  as  well  as 
the  fresh  water  supply  system  –  if  surface  water 
is  used  (Frehland  et  al.,  2020;  Yee  et  al.,  2021; 
Browne et al., 2011). Nanoplastics also leach from 
soil into rivers and oceans via natural erosion. The 
United Nations Environment Programme estimates 
that  4.18-12.7  million  tonnes  of  plastic  leached 
into the water systems in 2010 (Yee et al., 2021).

Aerial drone photo of latest technology sewage and sludge processing plant in small island of Psitalia or Psyttaleia, Piraeus, Attica, 
Greece (©Shutterstock, photo by Aerial-motion) 

[N]ovel technologies are highly demanded after tertiary treatment to 
eradicate small size particles (i.e., NPs, <0.1 μm) and fiber-like MPs  
(15–0.1 μm) from the final effluent to protect receiving water 
environment from plastic pollution 

Ali et al., 2021 

 
 
Box 8: Agricultural sources of micro- and nanoplastic pollution

31

On land, much microplastic pollution is associated 
with  agriculture  –  and  therefore  also  the  product 
of  microplastic  degradation,  nanoplastic.  Sources 
linked to the agricultural sector include:

 y Microplastics and nanoplastics in sewage 
sludge  –  a  large  proportion  of  ‘down  the 
drain’ plastics, where they are not removed by 
wastewater  treatment,  may  be  released  onto 
agricultural  soil  via  application  of  biosolids 
(ECHA,  2020).  One  study  estimated  the 
yearly amount of microplastics transferred to 
agricultural land from urban sources – including 
via sewage sludge – at between 125 and 850 
tonnes of microplastics per million inhabitants. 
This is equivalent to an annual input of 63 000 
–  430  000  tonnes  across  Europe  (Nizzetto, 
Futter and Langaas, 2016). Preliminary findings 
on the fate of PS nanoplastics found that 98% 
of nanoplastics were present in sewage sludge 
after  wastewater  treatment  (Frehland  et  al., 
2020).  

 y Microplastic  in  composted  domestic  and 
industrial waste – applied to soil as organic 
fertilisers  or  soil  conditioners  (Weithmann  et 
al., 2018). 

 y Residues 

from 

controlled-release 
fertilisers – plastic pellets in which fertilisers 
are contained persist in soil, known as ‘polymer 
encapsulation  systems’.  Anti  caking,  and  anti 
dust  polymer  additives,  are  also  a  source 
of  plastics  added  to  agricultural  fertilisers. 
However,  from  2021  within  Europe  the  EU 
fertilizing  products  regulation19  now  restricts 
the use of non-biodegradable plastics in most 
of  these  types  of  agricultural  fertilisers  – 
¬with a transition period of 5 years to comply. 

However,  controlled  release  fertilisers  are 
exempt  from  this  restriction.  Only  polymers 
meeting  biodegradability  requirements  of  the 
regulation  will  be  allowed  on  the  European 
market in most fertilisers from 2026.

 y Agricultural plastic mulches – as mulch film 
degrades  in  and  on  the  ground,  microplastics 
and  nanoplastics  form.  Soil  contamination  is 
around  467  kt  per  year  in  the  EU,  with  36% 
of  this  arriving  from  mulch  film  collection  – 
despite  mulch  being  only  12%  of  the  market 
(by  mass)20.  The  removal  of  plastic  mulch 
contaminated  soil  from  fields,  contributes  to 
the  loss  of  soil  organic  carbon  (SOC)  –  a  key 
component  of  soil  health.  Around  7  777  000 
tonnes of plastic mulch is used per year in the 
EU,  and  only  a  fraction  is  collected  after  use 
(EIP-AGRI, 2021). 

The  2018  standard  EN-17033  states  that 
biodegradable  mulch  films  should  have  a 
biodegradation threshold of 90% in 2 years, but 
since most crops only remain up to 6 months 
in the field – this is too long to wait before the 
ground  is  replanted.  Meanwhile,  only  some 
European countries, including France and Italy, 
have national standards for biodegradability of 
plastic mulch in arable soils. 

 y Discarded  plastics  from  greenhouses 
and  polytunnels  –  an  estimated  117  000 
tonnes  of  plastic  film,  primarily  LDPE,  is 
used  in  production  greenhouses,  and  56  300 
tonnes  in  smaller  tunnels  (EIP-AGRI,  2021). 
Fragments  from  degraded  film,  as  well  as 
mismanaged  or  abandoned  plastics,  end  up  
in the soil, rivers, and the sea.  

19  https://eur-lex.europa.eu/legal-content/EN/ALL/?uri=CELEX:32019R1009
20  EC (2021) Conventional and Biodegradable Plastics in Agriculture: https://environment.ec.europa.eu/system/files/2021-09/

Agricultural%20Plastics%20Final%20Report.pdf

32

Box 9: Nanoplastics from tyre wear

Tyre  wear  particles  (TWP)  are  produced  under 
normal  driving  conditions,  especially  during 
cornering,  braking,  accelerating,  and  other 
manoeuvres. In western Europe, emissions of TWP 
have  been  estimated  at  0.5–1.5  kilogrammes  of 
TWP  per  capita  per  year  (Kole  et  al.,  2017).  They 
may  contribute  up  to  10%  of  all  ocean  plastic 
(ibid.), and 40% of microplastics transported from 
rivers to seas (Siegfried et al., 2017). 

Composed  primarily  of  synthetic  and  natural 
rubber,  tyre  material 
is  strictly  speaking  an 
elastomer – belongs to the group of elastomers – 
flexible polymers, distinct from plastic – however, 
the particles emitted are often classified as micro- 
and  nanoplastics.  Emissions  are  typically  at  the 

micro  scale,  but  nano  scale  TWP  have  also  been 
detected (Gillibert et al., 2022).

TWP  typically  enter  the  environment  through 
rainfall  runoff  from  road  surfaces,  with  around 
75% estimated to enter roadside soils – considered 
a  sink  that  limits  the  further  spread  of  particles. 
However,  about  a  fifth  of  emissions  may  reach 
standing water (Baensch-Baltruschat et al., 2021). 
Where  runoff  enters  urban  sewage  systems, 
treatment  facilities,  such  as  soil  filters  that  can 
remove  nano-scale  particles  could  be  employed 
(ibid.).  Using  a  simulated  municipal  waste  water 
treatment plant, Mitrano et al. (2019) found more 
than  98%  of  nanoplastic  entering  the  facility 
remained in sewage sludge.

3.2 Degradation

Once released into air, soil and water, microplastics 
and  nanoplastics  will  be  subject  to  transport  and 
fragmentation,  and  some  types  to  biodegradation 
(Figure  7).  Abiotic  degradation  and 
processes 
transformation 
physical 
degradation,  photodegradation, 
thermo-oxidative 
degradation and hydrolysis (Yee et al., 2021). Physical 

processes 

include 

degradation  –  where  larger  plastics  fragment  into 
smaller pieces – occurs naturally through weathering. 
Such  processes  decompose  polymeric  structures, 
altering  their  mechanical  properties  and  increasing 
their  specific  surface  area,  resulting  in  enhanced 
potential 
reactions,  and 
interactions with microorganisms (Yee et al., 2021).

for  physical-chemical 

Table  3.  Types  /  pathways  of  plastic  particle  degradation,  from  Boyle  and  Örmeci  (2020).  Biotic  and  abiotic 
processes can occur simultaneously.

Abiotic degradation: mechanical degradation of plastics through weather and climate changes 
(e.g., freezing, thawing, pressure changes, water turbulence and damage by animal activities) – only 
morphological changes occur.

Thermal

Oxidation 

Breakdown of plastics 
at high temperatures; 
while often not 
hot enough in the 
environment, plastic 
production involves 
thermal treatment

Thermal or 
photo-induced; 
introduces oxygen 
to plastics and 
forms degradation-
promoting carbonyl 
and hydroxyl groups

Hydrolytic

Reaction of plastic 
compounds with 
water, resulting in 
changed molecular 
weight and reduced 
strength

Photo

Most damaging 
to environmental 
plastics; ultraviolet 
and visible light excite 
polymeric electrons 
and cause broken 
bonds

Biotic degradation: degradation resulting from organism activity – changes in chemical 
structure occur over time. Organisms secrete enzymes that break polymer chains and weaken 
some types of plastics.

 
33

Complete  degradation  returns  nanoplastic  back 
to  its  constituent  elements  –  largely  carbon  and 
hydrogen,  among  others.  Part  of  this  process 
involves breaking ionic and covalent bonds. Light, 
for example, can cause a lot of damage to plastic 
by increasing electron activity within these bonds, 
resulting  in  their  eventual  cleavage  (Boyle  and 
Örmeci, 2020). Abiotic processes can, in this way, 
break the long chain molecules in plastic polymers 
into  shorter  chains,  and  after  a  long  time  will 
eventually  convert  them  into  their  monomeric 
forms (Yee et al., 2021). 

Usually,  biodegradation  is  used  to  describe  the 
breakdown  of  material  by  micro-organisms,  and 
biotransformation  breakdown  occurring  within 
organisms,  e.g.  via  enzymatic  digestion.  When 
enzymes  break  up  plastic  polymer  chains,  it 

reduces  plastic’s  molecular  mass,  promoting 
further microbial degradation – enhanced by water 
and oxygen exposure. This gradually breaks down 
the  nanoplastic  polymer  molecule  into  shorter 
chains,  and  eventually  water-soluble  oligomers 
and  monomers,  that  mineralise,  and  can  be  used 
by  microorganisms  as  carbon  and  hydrogen 
sources  (Boyle  and  Örmeci,  2020).  Some  studies 
have shown that this can happen to biodegradable 
plastics  in  composting  situations  (Jakubowicz, 
2003), while a range of fungi have also been found 
to  potentially  degrade  different  petroleum-based 
plastics  –  an  application  that  could  be  further 
explored  in  the  context  of  remediation  (Sánchez, 
2020). It has also been shown that enzymes from 
environmentally  ubiquitous  fungi  can  break  down 
plastic additives, such as phthalates and bisphenol 
A (Carstens et al., 2020).

In this study the ability of … fungi to biotransform the PE representatives 
DBP and DEP, and the plastic precursor chemical BPA was demonstrated. 
… [M]etabolites in fungal cultures confirmed DBP biotransformation … The 
results of the present study imply an environmentally ubiquitous fungal 
potential for the biocatalytic breakdown of plastic additives

Carstens et al., 2020

Fig. 7. Degradation pathways of plastic materials in flowing water.  Source: Sarkar et al. (2020) 

 
34

Some  studies  have  suggested  a  degradation 
pathway within the digestive systems of organisms. 
Four days after Antarctic krill (Euphausia superba) 
ingested  virgin  polyethylene  (PE)  microplastics 
(particles  sized  31.5  µm),  measurements  showed 
they  had  reduced  in  size  by  78%  (7.1µm  ±  6.2 
SD)  (Dawson  et  al.,  2018).  The  authors  suggest 
the  bulk  of  PE  MP  breakdown  in  the  Krill  is  due 
to  the  stomach  and  gastric  mill  mechanically 
fragmenting  the  plastic  particles,  however,  they 
note, they cannot rule out that digestive enzymes 
also contributed to this breakdown (Dawson et al., 
2018).  Due  to  their  predominantly  herbi-  vorous 
diet, Antarctic krill have complex digestive enzymes 
with high activity. Another study found the larvae 
of  the  greater  wax  moth,  Galleria  melonella,  can 

also metabolise polyethylene (PE), possibly due to 
its chemical structure being similar to honeycomb 
–  the  larvae’s  main  food  (LeMoine  et  al.,  2020). 
The biotransformation phenomenon demonstrated 
in these studies could be investigated further, with 
other types of plastic (Piccardo, Renz and Terlizzi, 
2020). Recently, Beale et al. (2022) suggested that 
biotransformation via insects, such as black soldier 
fly  larva  (Hermetia  illucens),  and  mealworms 
(Tenebrio molitor), could also be investigated in the 
context  of  addressing  plastic  waste.  The  bulk  of 
plastic  maceration,  presumably  took  place  in  the 
stomach and gastric mill, which is responsible for 
mechanically  fragmenting  food  particles  under 
normal feeding conditions.

3.3 Environmental fate

Fig. 8. Schematic diagram showing sources and fate of plastics in the marine environment.  
Source: Gangadoo et al. (2020)

To  understand  the  potential  risks  of  nanoplastic 
particles  in  the  environment,  it  is  important  to 
ascertain  in  what  parts  of  the  environment  they 
collect  –  their  environmental  fate.  The  fate  of 
microplastics  is  better  understood  than  that  of 
nanoplastics  –  for  instance,  it  is  known  that  they 
can be washed into rivers during flooding – but there 

is still insufficient knowledge to reliably model the 
movement of either across environmental habitats. 
Existing chemical fate models, such as ‘SimpleBox21’ 
–  an  established  model  already  used  in  the  risk 
assessment  of  chemicals  –  may  be  modified  and 
applied to microplastics and nanoplastics (Kooi, et 
al., 2017). SimpleBox4Nano and NanoDUFLOW, are 

21  https://www.rivm.nl/en/soil-and-water/simplebox

35

two  examples  of  models  adapted  for  analysis  of 
nanoparticle fate (Meesters et al., 2014; Besseling 
et al., 2017).

Most research into nanoplastic fate has focused on 
pollution in the water column of bodies of water, and its 
effect on open water organisms, for example, pelagic 
fish  or  phytoplankton.  However,  over  time,  plastic 
particles tend to aggregate (join together with other 
particles) and sink. Several plastics are denser than 
water (PVC, PS, PET) and so sink quite rapidly, while 
others  like  low-density  polyethylene  (LDPE),  high-

density polyethylene (HDPE) and polypropylene (PP), 
mainly float in the water column, until being subject 
to aggregation or biofouling (formation of a biofilm of 
bacteria and organic matter) (Haegerbaeumer et al., 
2019). Relatively high concentrations of nanoplastics 
in  sediment,  compared  to  surrounding  water,  could 
pose  a  risk  to  benthic  (bottom  dwelling)  fauna. 
Ingestion of nanoplastics by benthic invertebrates is 
of particular concern, as they make up 90% of the 
biomass  eaten  by  fish,  with  implications  for  trophic 
transfer (ibid.).

Box 10: Do nanoplastics behave the same in freshwater and seawater?

In  order  to  predict  the  potential  distribution  and 
transport of nanoplastics in real world environments, 
it is important to investigate their different behaviours 
in  fresh-  and  seawater.  For  example,  increases  in 
particle  sizes  have  been  observed  in  studies  using 
engineered  nanoplastics 
in  seawater,  bacterial 
mediums and biological fluids (Gangadoo et al., 2020; 
Reynaud et al., 2022). Recently, Lee and Fang (2022) 
found  that  particles  exhibited  less  aggregation  in 

freshwater and stronger aggregation and deposition 
in  seawater  –  due  to  ionic  forces  –  increasing  with 
temperature.  Benthic  organisms  in  warmer  waters 
may  be  exposed  to  more  nanoplastics  than  in 
cooler  waters,  the  researchers  suggest.  Venel  et  al. 
(2021),  using  a  'lab-on-a  -chip',  showed  that  larger 
nanoplastics may aggregate and clump in estuaries.

Fig.  9.  Laboratory  studies  suggest  nanoplastic  particles  will  remain  more  dispersed  in  freshwater  than 
seawater.  Higher  temperatures  may  lead  to  increased  homoaggregation  and  deposition  in  sediment.  
Source: Lee and Fang (2022)

Most  studies  looking  at  nanoplastic  behaviour  in 
water  have  used  engineered  study  particles,  but 
the  behaviour  of  actual  (primary  or  secondary) 
nanoplastic  particles  may  be  quite  different.  For 
example,  rough-surfaced  and  irregularly  shaped 
particles are more likely to aggregate than smooth 
particles (e.g. Veclin et al., 2022), and coronas may 
form  on  particles  through  interaction  with  other 
substances present (see section 4.4).

Meanwhile, Schür et al. (2021) note that polystyrene 
microplastics became less toxic to Daphnia magna 
after  adsorbing  organic  matter  in  wastewater, 
inferring 
the 
environmental effects of pristine plastics – micro or 
nano – may not reflect real-world conditions.

that  experiments 

looking  at 

36

Fig.  10.  From  Brewer,  Dror  and  Berkowitz  (2020),  showing  the  aggregation  and  sedimentation  behaviour  of 
plastic  nanoparticles  (abbreviated  here  as  PNPs)  as  they  interact  with  inorganic  colloids,  metal  cations  and 
dissolved and particulate organic matter (DOM, POM)

3.4 Estimating nanoplastics in field samples

Current  technological  limitations  on  quantifying 
environmental  nanoplastics  mean 
that  most 
research on nanoplastics as pollutants has focused 
on laboratory evidence, however some recent studies 
have used novel techniques to detect particles in field 
samples (Gillibert et al., 2019; Shen et al. 2019; Cai 
et al., 2021a).

In  a  recent  review,  only  five  studies  reported  the 
detection  and  analysis  of  nanoplastics  in  real  field 
samples  (Cai  et  al.,  2021b),  though  new  data  is 

reliable  quantification 

being  published.  Nanoplastic  particles  are  easily 
destabilised,  so  many  can  be  lost  during  sample 
preparation,  making 
in 
samples  difficult.  Despite  not 
environmental 
being  able  to  estimate  the  exact  concentration  of 
nanoplastics  in  the  environment  at  present,  the 
expected  trend  is  that  they  will  increase  over  time: 
because of increased use in products, as a by-product 
of industry, and from the degradation of macro and 
microplastics  (Besseling  et  al.,  2019;  Koelmans, 
Besseling and Shim, 2015). 

Fig.  11.  Methodologies  for  detecting  and  quantifying  nanoplastics  in  environmental  samples  are  still  being 
developed. Source: Jakubowicz, Enebro and Yarahmadi (2021)

 
 
 
 
37

Nanoplastic  occurrence  in  marine  samples  has  so 
far only rarely been demonstrated. 

One  study  found  nanoplastics  present  in  ocean 
surface samples from the North Atlantic subtropical 
gyre (Ter Halle et al., 2017). The researchers detected 
nanosized  polymers  in  samples  taken  from  the 
ocean surface using pyrolysis gas chromatography-
mass spectrometry. Materić et al. (2022) quantified 
nanoplastics in surface water samples from Sweden 
and the Siberian Arctic tundra. They found four types 
of nanoplastic at a mean concentration of 563 µg/L 
in  the  Swedish  samples,  and  PVC  and  polystyrene 
nanoparticles were detected in Siberian samples, at 
a mean concentration of 51 µg/L.

Another  study  detected  PET  nanoplastic  in  Alpine 
snow  samples,  semi-quantifying  its  concentration 
in surface snow as 18.5 ± 1.5 ng/ml of filtered snow 
(Materić  et  al.,  2020).  The  authors  suggest  this 
observation  shows  airborne  nanoplastic  particles, 
largely PET, are being deposited on the surface of 
the snow. Another analysis of Alpine snow in Austria 
found 46.5ng/ml of melted surface snow, including 
polypropylene as well as PET (Materić et al., 2021). 
Drawing on air transport modelling, the researchers 
inferred  an  average  deposition  rate  of  42  kg  of 
nanoplastic  per  square  kilometre  each  year  in  this 
remote location. The particles mostly originate from 
urban Europe, they note. 

The presence of nanoplastics in high-altitude snow indicates 
airborne transport of plastic pollution with environmental and health 
consequences yet to be understood 

Materić et al., 2020

Another  study  demonstrated  the  use  of  klarite 
substrates (gold cavities that can hold nanoplastic 
particles),  in  conjunction  with  surface-enhanced 
Raman  spectroscopy,  to  collect  and  analyse 
nanoplastics  in  ambient  airborne  samples  (Xu  et 
al., 2020). Klarite may allow detection of particles 
as  small  as  360  nm.  Airborne  particle  sampling 
is  important  due  to  the  potentially  high  levels  of 
exposure via inhalation in indoor environments. 

In  contrast,  many  studies  looking  to  measure 
in  the  environment  have  been 
microplastics 
published. Globally, the highest reported microplastic 
concentrations  in  the  aquatic  environment  are 
102  particles/L  water,  and  1529  particles  per 
kg  sediment,  with  a  large  range  dependent  on 
region and type of habitat i.e. freshwater, marine, 
estuarine  (beach  is  higher  than  this  quoted 
maximum).  Fragmentation  of  microplastic  to 
nanoplastic has been measured in the laboratory 
and  shown  in  the  environment  as  well  (Gigault 
et al., 2016; Gillibert et al., 2019). It is not known 
how fast, or to what extent nanofragmentation of 
plastics  will  occur  in  the  environment;  however, 
based on mass conservation principles, if spherical 
microplastic particles with a size of 0.1 µm to 5 mm 
degraded  into  100  nm  nanoplastic  particles,  this 

would produce nanoplastic particle concentrations 
that  are  more  than  1014  times  higher,  than  the 
current  microplastic  concentrations 
(Besseling 
et  al.,  2019).  This  is  conservative  as  it  doesn’t 
include the ongoing degradation of macroplastics 
to  microplastics,  and  ultimately  nanoplastics.  It 
is  estimated,  via  model-based  studies,  that  this 
process  of  nanofragmentation  of  plastics  would 
occur  over  several  hundreds  of  years  (Koelmans, 
Besseling and Shim, 2015). 

Given  the  paucity  of  data  on  nanoplastic  levels 
in  environmental  samples,  potential  distribution 
in  air,  aquatic  and  terrestrial  environments  can 
only  be  inferred  from  data  on  microplastics  and 
modelling  studies.  For  example,  the  highest 
reported ranges of microplastics are in near shore 
or estuarine areas, compared with open ocean or 
freshwater  systems  (Besseling  et  al.,  2019).  The 
authors suggest this could be due to accumulation 
of  microplastics  (and  presumably  nanoplastics), 
due  to  input  from  rivers  and  beaches,  as  well  as 
backwashing from marine currents, and nearshore 
hydrodynamics 
trapping/fouling  microplastics. 
However, as it is easier to sample near shore areas, 
the  higher  concentrations  reported  may  also  be 
influenced by the sampling effort.

38

Kooi  et  al.  (2017)  used  a  model  to  calculate  the 
average,  background  predicted,  environmental 
steady  state  concentrations  (PECs)  of  plastic 
different 
particles 
compartments  in  the  Rhine  Catchment  in  Europe. 
This calculation assumed a yearly emission of 20 
kilotons,  with  50%  emitted  to  water  and  50%  to 

different 

sizes, 

of 

in 

soil – the results can be seen in Table 4 below. The 
nanoplastic-sized  particles  in  the  first  column  of 
the table, can be seen to mirror the trend observed 
for  larger  microplastic  particles  up  to  10µm  in 
levels of accumulation – most being found in the 
soil, fewer in water, and fewer still in sediment. 

Table  4.  Distribution  of  plastic  particles  of  different  sizes  over  the  soil,  water  and  aquatic  sediment 
compartments,  as  predicted  by  the  multi-media  model  SB4N.  PECs  are  based  on  a  yearly  emission  of  20  Kt, 
and  assumes  a  fouled  plastic  density  of  1100kg/m3,  negligible  degradation  and  fragmentation  due  to  short 
particle  residence  time  in  the  system,  along  with  an  attachment  efficiency  for  heteroaggregation  of  0.01.  
(Kooi et al., 2017)

Particle size

0.1 µm 

1 µm

10 µm

100 µm

1 mm

soil (log μg/m3)

water (log μg/m3)

sediment (log μg/m3)

6.43

5.45

1.52

6.38

5.44

2.41

6.17

5.39

4.42

4.57

4.89

6.07

2.62

3.08

6.26

3.5 Nanoplastics in aquatic organisms

There is  a plethora of experimental and  field data 
showing  that  microplastics  can  be  ingested  by  an 
array  of  species  from  different  taxonomic  groups, 
and  occupying  various  ecological  niches,  and 
positions along food chains. Ingestion has currently 
been documented in around 220 species (Lusher et 
al.,  2017).  Although  fewer  studies  exist  regarding 
nanoplastics,  interaction  with  aquatic  plants  and 
ingestion by animal species has been demonstrated 
–  including  marine  birds,  fish,  mammals,  mussels, 
crustaceans,  mussels  and  zooplankton    (Larue  et 
al.,  2021;  Chae  and  An,  2018;  Barría  et  al.,  2020; 
Boyle and Örmeci, 2020). Uptake via gills provides 
an additional exposure route for fish to nanoplastic 
particles. Nanoplastics may induce adverse effects 
in aquatic organisms, as their small size means they 
can  be  ingested,  and  also  pass  through  biological 
membranes – unlike microplastics, which are often 
too large to pass through (Boyle and Örmeci, 2020). 
The  small  size  of  nanoplastics,  means  they  have 
the potential to penetrate any part of an organism, 
causing undesirable effects due to the presence of 
the particle itself (Boyle and Örmeci, 2020; Hartman 
et al., 2017). Chapter 4 discusses ecotoxicity in more 
detail. As noted, the infinite combination of plastic 

characteristics and potential interactions throughout 
the  marine  ecosystem,  coupled  with  a  paucity  of 
techniques  to  quantify  amounts  of  nanoplastics 
ingested,  are  challenges,  making  risk  analyses 
of  nanoplastics  for  marine  organisms  difficult 
(Galloway,  Cole  and  Lewis  2017;  EASAC,  2020). 
Meanwhile,  there  is  a  risk  that  bioaccumulation  of 
nanoplastics could occur after ingestion by aquatic 
organisms,  including  fish  and  marine  mammals. 
Organisms  at  higher  trophic  levels  (higher  in  the 
food chain) may also ingest nanoplastics present in 
organisms they consume. 

Studies have shown that microplastics ingested are 
frequently egested (excreted), and the ECHA (2019) 
report suggests the accumulation of microplastics 
in  fish,  for  example,  is  low  as  a  result.  However, 
invertebrate  species  such  as  scallops  have  been 
found to ingest microplastics and nanoplastics, with 
some particle sizes remaining in the organism after 
a  protracted  time,  and  evidence  of  translocation 
within the organism has also been highlighted (Al-
Sid-Cheikh  et  al.,  2018).  Understanding  to  what 
extent ingesting small plastic particles differs from 
ingesting natural particles of sediment, or natural 

 
39

organic  material  of  similar  size,  is  important,  but 
there  are  few  comparative  studies.  One  study  on 
differing  responses  to  microplastics  and  kaolin 
clay  in  Daphnia  magna  found  that  secondary 
microplastics have a greater capacity to negatively 
affect  feeding  behaviours  compared  to  naturally 
occurring particles of a similar size (Ogonowski et 
al., 2016); the same may be true for nanoplastics. 
Likewise,  a  study  comparing  the  effects  of 
ingestion of a bio-based microplastic, a petroleum-
based  microplastic  and  natural  silica  particles 

as  a  control,  found  that  both  types  of  plastic 
had  detrimental  effects  on  aquatic  invertebrate 
species,  Gammarus  fossarum  (Straub,  Hirsch  and 
Burkhardt-Holm,  2017).  Another  study  comparing 
the  effects  of  microplastics  PVC  and  PMMA  with 
natural  microparticles  on  mussels  (Semimytilus 
algosus), found that significantly different negative 
reactions    to  microplastics  only  occurred  at  high 
concentrations that are unlikely to be found in the 
environment (Barkhau et al., 2022). 

… mussels exposed to PVC … microplastics produced over 40% less 
byssus than those exposed to natural microparticles. This suggests that 
mussels react differently to natural microparticles and to microplastics, 
but only at high particle loads that exceed current environmental 
microplastic concentrations by orders of magnitude

Barkhau et al., 2022

Because  of  their  ability  to  cross  biological 
membranes,  it  can  be  expected  that  nano-sized 
particles  might  confer  worse,  or  different,  toxic 
effects  at  lower  concentrations..  The  very  size  of 
nanoparticles  –  of  plastic  or  any  material  –  may 
make them more hazardous to aquatic organisms 
than larger particles. Some organisms will excrete 
indigestible  nanoplastic  particles;  others  may 

absorb  smaller  nanoplastic  particles  through 
membranes,  leading  to  potentially  toxic  effects 
in  various  organs.  Much  concern  is  also  focused 
on  the  toxic  additives  carried  by  plastic  particles, 
which  could  leach  into  the  bodies  of  organisms 
that  ingest  them  (EASAC,  2020).  Mechanisms  of 
toxicity are discussed further in Chapter 4.

Fig. 12. Oysters appear to select which particles to ingest based partly on their shape. These microscope images 
show polyester fibres (ca. 65-260 μm long × 16 μm wide, red arrows) and microspheres (10-20 μm diameter, yellow 
arrows) on the gills of Crassostrea virginica. Only the microspheres are transported to the dorsal tract, indicating 
their selection for ingestion. Scale bars ≈ 200 μm. Source: Ward, Rosa and Shumway (2019)

40

3.5.1 Translocation of nanoplastics in aquatic vertebrates

‘Translocation’  refers  to  the  passage  of  particles 
across  biological  membranes,  where  they  end  up 
in  the  tissues  of  organisms.  A  range  of  research 
indicates that nanoplastics do translocate in fish – a 
finding often correlated with negative health effects 
(Barría et al., 2020). In the laboratory, for example, 
researchers  have  demonstrated 
translocation 
of  nanoplastics  across  salmon  intestines  using 

palladium-doped  polystyrene  (Clark  et  al.,  2022); 
translocation  to  liver  and  kidney  of  fathead 
minnow (Pimephales promelas) following exposure 
to  fluorescent-labelled  polystyrene  nanoparticles 
via  diet  and  injection  (Elizalde-Velázquez  et  al., 
2020); and extensive internalisation of nano-sized 
polystyrene into zebrafish cell cultures (Sendra et 
al., 2021). 

Fig. 13. Experiments with polystyrene nanoparticles have shown that they are able to reach fish organs, with 
detrimental health effects. Source: Elizalde-Velázquez et al. (2020)

Polystyrene  nanoparticles  were  also  studied  by 
Sökmen, et al. (2020), who found they reached the 
brains  of  zebrafish  larvae.  Mattson  et  al  (2015) 
noted  that  brains  of  fish  exposed  to  nanoplastics 
were heavier, whiter, and fluffier in appearance than 
controls,  possibly  due  to  the  lipophilic  nature  of 
polystyrene chains being attracted to the lipid-rich 
brain tissue. A study using latex particles of <50nm 
found  these  particles  accumulated  throughout 
the  bodies  of  Japanese  rice  fish  (Oryzias  latipes) 
–  in  testes,  liver,  blood  and  brain  –  also  causing  a 
decrease in survival of affected fish embryos (Boyle 
and Örmeci, 2020). In another study using diet fed 
exposure  of  fish  to  PS  nanoplastics,  nanoparticles 
were identified in the yolk sac, liver and pancreas of 
larvae and embryos after maternal dietary exposure 
–suggesting  maternal  to  offspring  transfer  in  fish 
(Pitt et al., 2018). 

The  breaching  of  the  blood-brain  barrier  in  zebrafish 
and Japanese rice fish is particularly concerning for both 
human and animal health. Importantly, however, many 
translocation and toxicity studies use fluorescent-marked 
particles; some research has questioned the validity of 
this  method,  suggesting  it  may  lead  to  inaccuracies  in 
determining  translocation  of  nanoplastic  in  organisms 
(Schür et al., 2019; Catarino, Frutos & Henry, 2019). That 
is,  fluorescent  compounds  themselves  may  leach  from 
marked  particles  and  accumulate  in  model  organisms, 
so  observations  must  account  for  this  (Figure  14). 
Radio-labelled  or  trace  metal  doped  particles  offer  an 
alternative  option  (e.g.  Redondo-Hasselerharm  et  al., 
2021).  Additionally,  a  large  proportion  of  studies  use 
polystyrene particles, which may not well represent the 
behaviour of other types of nanoplastic within organisms; 
nor is it the most commony found microplastic found in 
the environment (and by extension, nanoplastic). 

41

translocation 

in  fish  sampled 
Investigating 
from  the  field,  meanwhile,  faces  the  challenge 
of  separating  and  quantifying 
(unmarked) 
nanoparticles  in  biological  tissue,  while  avoiding 
destruction of these target particles during sample 
processing.  Studies  that  detect  translocation 
of  nanoplastics  in  fish  collected  from  fresh 
and  seawater  environments  are  lacking.  One 
researcher  attempted  to  retrieve  nanoplastics 
from dissected fish from the Aegean Sea, but was 
unable to ascertain whether nano-sized particles 
filtered out were plastics (Gimskog, 2019).

A  similar  study  with  a  commercially  important 
scallop  species  eaten  by  humans,  Pecten 
maximus  (great  or  king  scallop),  found  that  at 
environmentally relevant levels (<15 µg per litre of 
water), nanoplastic PS particles of 24 nm and 250 
nm were taken up rapidly into the scallop body (Al-
Sid-Cheikh  et  al.,  2018).  In  this  study,  Carbon-14 
isotopelabelled  PS  particleswere  used,  allowing 
precise tracing of the administered particles within 
the  scallops.  The  larger  250  nm  nanoplastics 
accumulated in the intestine, whereas the 24 nm 
particles  were  distributed  throughout  the  whole 
body  –  suggesting  translocation  across  epithelial 
membranes  of  the  scallop.  After  14  days  the  24 
nm  particles  were  no  longer  detected,  whereas 
the  250  nm  particles  were  still  present  after  48 
days  –  indicating  differential  removal  of  different 
particle  sizes.  Chronic  exposure  studies  would  be 
beneficial to gain a greater understanding of this 
mechanism.  Modelling  suggested  that  after  300 
days  of  environmental  exposure  to  nanoplastics, 
the scallop body tissues would reach equilibrium – 
with concentrations below 2.7 mg/g (ibid.).

Fig. 14. Recent work has highlighted a methodological 
problem when using fluorescent-marked nanoparticles 
to  trace  nanoplastics  in  fish  tissue.  Source:  Catarino, 
Frutos & Henry (2019) 

3.5.2 Translocation in aquatic invertebrates

As  is  the  case  for  the  fish  studies,  translocation 
in  aquatic  invertebrates  such  as  shellfish  are 
also  mostly  carried  out  with  polystyrene  (PS) 
nanoparticles. For example, an experiment with the 
Mediterranean  mussel,  Mytilus  galloprovincialis, 
demonstrated the translocation of PS nanoplastics 
into the hemoctyes (immune system cells) (Sendra 
et  al.,  2020).  The  smallest  PS  nanoplastics 
tested  were  detected  in  the  digestive  gland  and 
muscle,  with  swift,  size-dependent,  translocation 
to  the  haemolymph  (the  invertebrate  equivalent 
of  blood),  after  just  3  hours  of  exposure.  The 
exposed  hemocytes  suffered  functional  losses 
in  motility,  cell  death,  reactive  oxygen  species 
(ROS)  and  loss  of  phagocytic  capacity  (ability  to 
surround and destroy micro-organisms and foreign 
material).  However,  the  hemocytes  were  resilient 
when  infected  with  bacteria  after  this  exposure, 
recovering 
their  phagocytic  capacity,  despite 
expression  of  the  mussels’  antimicrobial  peptide, 
myticin C, being lowered. 

42

Fig.  15.  Pecten  maximus,  scallop  exposed  to  14C-radiolabeled  nanopolystyrene  at  predicted  environmental 
concentrations of (<15 μg/L). Uptake was rapid, shown by autoradiograph, and greater for 24 nm than for 250 nm 
particles. After 6 hours, nanoplastics of 250 nm accumulated in the intestine, while 24 nm particles were dispersed 
throughout the whole-body, possibly indicating some translocation across epithelial membranes. Red, yellow and 
green in the right hand x ray image, shows the presence of 24nm radio labelled PS particles, with red indicating 
lower concentrations, yellow medium, and green higher concentrations.  Source: Al–Sid–Cheikh et al. (2018)

research 

Some 
that  nanoplastic 
suggests 
particles  may  have  transgenerational  effects  via 
translocation to reproductive organs. For example, 
toxicity  was  observed  in  the  nematode  species 
Caenorhabditis  elegans 
following  prolonged 
exposure  to  PS  nanoplastics  at  concentrations 
higher than 10 µg/L (Zhao et.al., 2017).  When the 
concentration  was  increased  to  over  100  µg/L, 
transgenerational  effects  were  observed;  the 
authors suggested this was due to translocation of 
PS nanoplastics into organs such as the gonad and 
possible transfer to the next generation. The effect 

was not found to be due to leachates. In another 
study, Daphnids fed algae which had been exposed 
to  iron-tagged  270  nm  diameter  PS  were  found 
to contain these nanoplastic particles, suggesting 
trophic  transfer.  These  daphnids  took 
longer 
to  produce  their  first  offspring,  and  had  lower 
numbers.  In  the  same  study,  the  next  generation 
of  daphnids  from  parents  exposed  to  europium 
labelled  PS  nanoparticles  contained  a  traceable 
amount of europium – implying it was transferred 
from parents to offspring (Monikh et al., 2020). 

3.5.3 Trophic transfer of nanoplastics 

Given the confirmed presence of microplastics (and 
presumably nanoplastics) in a range of taxa, it has 
been suggested that trophic transfer of micro- and 
nano-plastics  could  occur,  through  aquatic  and 
terrestrial  food  chains.  Studies  have  confirmed 
transfer  of  microplastics  in  the  laboratory,  but  it 
is  thought  that  microplastics  may  be  subject  to 
significant gut clearance in fish (Güven et al. 2017; 
ECHA,  2020)  and  bivalves,  for  example,  which 
also  may  also  preferentially  capture  particles 
over  1um  (Ward,  Rosa  and  Shumway,  2019). 
However, when nanosized particles are ingested by 

organisms,  they  may  cross  biological  membranes, 
increasing the likelihood of uptake and potential for 
bioaccumulation, plus subsequent transfer through 
the food-chain (Besseling et al., 2014; Larue et al., 
2021; EASAC, 2020). 

Mattsson  et  al.  (2015)  administered  24  nm  and 
27  nm  PS  nanoparticles  to  carp  fish  via  algae 
and  zooplankton.  The 
researchers  observed 
severe  effects  on  feeding  and  shoaling  of  fish, 
demonstrating  the  possible  uptake  of  particles 
through  a  food  chain  affecting  behaviour  of  the 

 
43

Further studies using levels of nanoplastics found in 
the natural environment, i.e. at lower concentrations 
over a longer period of time, are warranted where 
there  is  potential  for  cumulative  build-up  in  the 
bodies  of  aquatic  vertebrates.  More  research 
taking  multiple  levels  of  trophic  transfer  into 
account  and  mixtures  of  nanoplastics  would  also 
improve  understanding  (Latchere  et  al.,  2021;  Zhu 
et al, 2021). A novel study by He et al. (2022), for 
example,  tracked  palladium-doped  nanoplastics  in 
a constructed freshwater ecosystem with zebrafish, 
three invertebrates and two aquatic plants, over 49 
days.  Results  showed  that  chronic  exposure  led  to 
more uptake than short duration exposures, but no 
biomagnification effect was identified – consumers 
higher in the food chain had not absorbed relatively 
more nanoplastics than organisms further down the 
chain. Potential bioaccumulation has been noted in 
studies  of  environmental  microplastics,  however: 
microplastics  adhered  to  seagrass  blades  in  a 
remote  California  reserve  were  found  at  a  higher 
density per gram in a herbivorous snail present at 
the  site,  than  on  the  two  species  of  sea  grasses 
examined (Saley et al., 2019). 

At present there is a paucity of data on the presence 
of phthalates in marine mammals, but also on how 
nanoplastics  might  contribute  to  the  presence 
of  these  and  other  plastic  additives  (Panti  et  al., 
2019). However, as microplastics have been shown 
to degrade to nanoplastics within the environment, 
and within organisms, areas with high microplastic 
concentrations  therefore  may  also  contain  high 
concentrations of nanoplastics. Although the source 
of  the  plasticisers  detected  in  marine  mammals, 
has not been definitively linked to any particular size 
of  plastic,  the  presence  of  plastic  pollution  of  any 
size in the ocean is a principal source of phthalates. 

top  consumer.  Nano-sized  PS  particles  have  been 
shown experimentally to transfer up to four trophic 
levels  within  an  aquatic  food  chain,  from  algae  to 
end-consumer fish species (Chae et al., 2018). In this 
experiment, the nanoplastics moved easily through 
the food chain with negative health impacts for the 
fish species exposed, which demonstrated lowered 
levels of activity and tissue changes in their livers. In 
addition, juveniles were found to have nanoplastics 
present  in  their  yolk  sacs.  However,  the  exposure 
dosages used in the study were artificially high, and 
fluorescence  as  a  marker  can  lead  to  inaccuracies 
(Schür  et  al.,  2019),  therefore  general  conclusions 
cannot be drawn from these findings. 

A  more  recent  study  found  that  algae  accumulate 
nanoscale plastic debris (NPD), which is transferred 
to  zooplankton  –  with  greater  uptake  and  trophic 
transfer  of  smaller  NPD  (270  nm),  compared  with 
larger  NPD  (640  nm)  (Monikh  et  al.,  2020).  In  this 
study, iron and europium were used as a label for 
the  nanoscale  PS,  rather  than  fluorescence.  The 
smaller  sized  NPD  were  found  to  detrimentally 
affect  reproduction  in  the  zooplankton  and  to  be 
passed on to the young.

3.5.4 Nanoplastics and marine mammals 

Nanoplastic  presence  in  the  ocean  is  of  concern 
because of possible toxicological impacts on aquatic 
organisms.  With  their  potential  to  bioaccumulate 
and be passed along food chains, conservationists 
fear  the  impact  this  might  have  on  predators 
and  large  filter  feeders,  such  as  polar  bears  and 
whales  (Routti  et  al.,  2021).  Plastic  additives  may 
add to this risk. Phthalates are a specific group of 
compounds of concern since these plasticisers are 
not chemically bound to the polymers of the plastic 
and may leach out of micro- or nanoplastic particles. 
These  can  then  enter  organisms,  or  be  ingested 
within plastic particles and then leach out internally 
(ibid.; Hahladakis and lacovidou, 2018;  Paluselli et 
al.,  2018).  Detection  of  phthalates  in  filter-feeding 
whales  from  the  Mediterranean  Sea  and  Sea  of 
Cortez has been correlated with high concentrations 
of microplastics in these areas (Fossi et al., 2016). 

44

Box 11: Gathering data on micro- and nanoplastics in cetaceans

b) Analysis  of  the  levels  of  plastic  additives 
(biomonitoring), as a proxy for ingestion: The 
levels  of  plastic  additives  (such  as  phthalates  or 
PBDEs) and associated Persistent Bioaccumulative 
and Toxic (PBT) compounds  –  in free ranging and 
stranded cetaceans  –  can aid in quantifying the 
exposure to marine plastic pollution;

c) Analysis of biomarker responses: Biological 
responses,  up  to  a  few  hours  after  death,  can  be 
used  to  detect  the  potential  toxicological  effect, 
related  to  PBT,  and  plastic  additives  from  plastic 
ingestion (Panti et al., 2019).

A  threefold  approach  to  assessing  the  impact  of 
marine  litter  on  cetaceans  has  been  suggested 
(Panti et al., 2019). This is to address the multiple 
potential,  physical,  and  ecotoxicological  effects 
of  marine  litter  interactions,  from  macro-,  micro- 
and  nanoplastics.  Data  on  rate  of  ingestion,  and 
sublethal  short-term  and  long-term  effects  of 
marine plastic litter, can be investigated using the 
threefold  approach  –  with  each  method  applied 
independently, or simultaneously, in either stranded 
or  free  ranging  animals  (Panti  et  al.,  2019).  The 
threefold approach consists of:

the  gastro-intestinal  content 

a)  Analysis  of  gastro-intestinal  content: 
Analysis  of 
in 
stranded  cetaceans,  to  determine  the  occurrence 
and rate of marine plastic litter ingestion and any 
associated pathology;

3.5.5 Nanoplastics, aquatic plants and algae

Positively  charged  plastic  particles  have  been 
shown to attach to the outside of phytoplankton. As 
these organisms are the first level in aquatic food 
chains, adhered plastic particles could be passed up 
to the next trophic level. Green algae, for example, 
have been shown to transfer plastic particles to the 
crustacean Daphnia magna, and to two species of 
fish. Nanoparticles of PS have a higher toxicity to D. 
magna (6x higher mortality), when they are exposed 
through  their  diet  –  ingesting  the  nanoplastics, 
rather than encountering them in the water column 
(Besseling et al., 2014). However, weathering usually 
induces an overall negative particle surface charge 
on plastic particles, which has been found to make 
them  less  likely  to  attach  to  algae,  as  cellulose  in 
algae also bears a negative charge.

also 

affects 

pollution 

aquatic 
Nanoplastic 
photosynthetic  organisms  such  as  seaweed 
(macroalgae)  and phytoplankton (microalgae). This 
merits investigation, as phytoplankton are not only 
the foundation of the aquatic food web, but are also 
responsible for fixing 45% of the carbon dioxide in 
the air, and producing about half the world’s oxygen. 
Aquatic  plants  (macrophytes)  also  offer  important 
ecosystem services.

Larue  et  al.  (2021)  conducted  a  review  of  studies 
on  the  impacts  of  nanoplastics  and  microplastics 
on  aquatic  plants.  Most  of  the  articles  focused  on 
nanoplastics, with 90% finding that they can have 
toxic  effects.  The  studies  noted  that  most  toxic 
effects  were  seen  with  smaller  size  particles  of 
plastics,  below  1  µm.  Leached  chemical  additives 
were also implicated in toxicity – and plasticisers such 
as phthalate esters have been shown to accumulate 
in some plants, resulting in exposure for people that 
eat  them.  The  toxicity  mechanism  for  nanoplastic 
particles  acting  on  phytoplankton  is  dominated  by 
the  chemical  effect,  i.e.  the  formation  of  reactive 
oxygen species and decreased photosynthesis; the 
converse  is  true  for  bigger  microplastic  particles 
where toxicity is driven by physical effects (ibid.).

45

Box 12: Factors influencing toxicity to phytoplankton

After  phytoplankton  exposure  to  nanoplastics, 
expression  of  genes  related  to  oxidative  stress 
regulation,  photosynthesis,  fatty  acid  synthesis, 
and  cell  aggregation  can  be  altered.  Some 
microalgae  species  appear  more  sensitive  than 
others (Larue et. al., 2021). 

Other factors that influence toxicity of nanoplastics 
to phytoplankton include:

 y Concentration

Larue  et  al.  (2021)  reported  that  environmental 
concentrations  of  PS  nanoparticles  at  which 
half  of  a  phytoplankton  population  is  affected 
ranged  between  0.5  and  13mg  per  litre.  The 
highest  reported  plastic  concentration  in  the 
aquatic  environment  so  far  is  1.56  mg  per  litre 
(Lasee  et  al.,  2017),  suggesting  plastic  pollution 
poses  a  risk  to  phytoplankton  in  these  highly 
contaminated aquatic regions. Few data expressing 
environmental  concentrations  in  mass  per  litre 
are  currently  available  (Larue  et  al.,  2021)  –  an 
important knowledge gap to address to assess risk 
to aquatic life. 

 y Type of plastic

PVC  has  often  been  found  to  be  more  toxic  than 
other  types  of  microplastic,  such  as  polystyrene, 
PP and PET (Zhu et al., 2019; Capolupo et al., 2020). 
This is likely to be due to the higher amount of toxic 
chemical additives released from this polymer. 

 y Aging of particles

Most  studies  on  the  toxicity  of  nanoplastics  (and 
microplastics) to phytoplankton and other species 
are  conducted  in  a  lab,  using  pristine  plastic 
model  particles  –  yet  it  has  been  shown  that 
aged  plastic  particles  taken  from  the  ocean  can 
have  different  effects.  In  a  study  examining  the 
impact  of  pristine  PVC  versus  environmentally 
aged  PS,  it  was  found  that  the  aged  plastic  was 
more  toxic  to  phytoplankton  Chlorella  vulgaris.  A 
similar  effect  was  noticed  using  ocean-aged  PE 
versus  pristine  PE  on  microalgae  Scenedesmus 
subspicatus  –  perhaps  related  to  molecular-level 
changes  in  the  structure  of  the  aged  PVC,  which 
conferred increased hydroxyl and aromatic groups. 
In  a  similar  study  with  PE  particles,  aged  plastic 
released more additives, meaning it was more toxic 
to  cyanobacteria  than  the  non-aged  PE  particle 
leachate (Larue et al., 2021).

 y Chemical combinations

indicated 

A  plethora  of  studies  have 
that 
microalgae  are  more  susceptible  to  toxic  effects 
from  nanoplastics,  and  other  ecotoxic  chemicals, 
when exposed to them in combination (e.g. Zhu et 
al., 2019). 

The  ability  of  microalgae  to  absorb  micro-  and 
nanoplastics has also been highlighted as a potential 
remediation and wastewater treatment tool (e.g. Manzi 
et al., 2022). Uptake of nanoplastics into macrophytes 
– aquatic plants such as duckweed (Lemna spp.) and 
water  hyacinth  (Eichhornia  crassipes)  –  meanwhile, 
has only recently been documented. 

In  one  study,  researchers  compared  the  effects  of 
positively  and  negatively  charged  PS  nanoplastics 
on  duckweed  at  different  concentrations  (Xiao  et  al., 
2022a). Uptake into plant tissues was observed after 
three days at concentrations of 10-50 µg/ml, together 
with adverse effects such as reduced growth. However, 
particles  did  not  appear  to  be  taken  up  after  seven 
days of exposure at concentrations lower than 0.1 µg/
ml,  which  the  researchers  say  are  environmentally 

relevant. Although there were no obvious effects on the 
plants  at  these  concentrations,  a  further  experiment 
exposing plants to particles at 0.015 µg/ml suggested 
that chronic exposure could induce genetic changes.

Other  studies  have  also  detected  micro-  and 
nanoplastics  on  the  root  surface  of  duckweed,  but 
no  evidence  of  internalisation  (Dovidat  et  al.,  2020). 
Nevertheless,  attached  particles  may  be  consumed 
by  organisms  that  feed  on  these  plants.  In  the 
Mediterranean  Sea,  for  instance,  microplastic  fibres 
and beads have been found adhered to the sea grass, 
Posidonia oceanica (Neptune grass), which forms balls 
of  fibrous  material  that  wash  up  on  sandy  beaches. 
Of the detritivorous fauna feeding on these, such as 
snails, researchers found just under 30% had plastic 
fibres in their digestive tract (Larue et al., 2021). 

46

Fig. 16. Seagrass ball on a Mediterranean beach. Organisms that feed on this aquatic plant may also be ingesting 
nanoplastic adhered to it. Credit: Ezu

3.6 Terrestrial habitats

Fig. 17. Potential impacts of plastic contamination: soil biogeochemistry; ingestion by birds; reduction 
in growth of earthworms; lethal toxicity to fungi; mammal lung inflammation and broad cytotoxicity of 
nanoplastics. Source: de Souza Machado et al. (2018)

 
47

Annual plastic production currently exceeds 380 million tons ... [and] 
it is estimated that roughly 32% of plastic waste might find its first 
receptacle in soils or continental aquatic ecosystems …  
[A]pproximately 4 977 million tons have accumulated in landfills  
and the natural environment 

Public attention on plastics has largely focused on 
the marine environment, perhaps in part due to the 
‘Blue  Planet  effect’,  referring  to  the  documentary 
series of the same name (Dunn, Mills and Veríssimo, 
2020). However, most marine pollution comes from 
land;  levels  of  microplastic  contamination  within 
terrestrial soil are much higher than in the ocean – 
some 4 to 23 times larger (de Souza Machado et al., 
2018; EASAC, 2020). 

There  is  a  growing  body  of  research  showing  that 
both micro- and nanoplastics interact with terrestrial 
organisms  from  plants,  fungi  and  earthworms  to 
birds,  rabbits  and  rats  –  many  of  which  provide 
essential  ecosystem  services  (de  Souza  Machado 
et  al.,  2018).  The  widespread  presence  and 
persistence of micro- and nanoplastics in the global 
environment, and their interactions with terrestrial 
and  freshwater  biota,  mean  these  small  plastic 
particles are viewed as an emerging global threat. 

3.6.1 Soil and earthworms

Studies  quantifying  the  amount  of  nanoplastics 
in soil are scarce, probably due to a lack of robust 
analytical  techniques.  However,  studies  addressing 
microplastics, which degrade to nanoplastics, have 
found  greatly  varying  concentrations  in  terrestrial 
soil  at  different  sites,  ranging  from  under  10  mg/
kg of soil to between 55.5 and 67 500 mg/kg (Büks 
and  Kaupenjohann,  2020;  Larue  et  al.,  2021).  A 
range of studies have also measured concentrations 
in  terms  of  numbers  of  items  (of  microplastic)  in 
dry  soil.  Büks  and  Kaupenjohann  (2020)  reviewed 
23  studies  on  microplastic  contamination  in  soils 
and found that median concentrations from studies 
conducted in Europe amounted to 2 914 items/kg, 
which is twice as high as in Chinese studies. 

that 
Researchers  have  expressed 
microplastics  could  migrate  through  soil  and 

concern 

Source: de Souza Machado et al., 2018

Further research is needed on the fate and effects 
of nanoplastics on land and in freshwater habitats 
(de Souza Machado et al., 2018; Nizzetto, Futter and 
Langaas, 2016; EASAC, 2020).

Nanoplastics are anticipated to be introduced to 
soils as a result of “landfill leachate, agricultural 
mulches,  application  of  wastewater  biosolids 
land,  or  by  direct  releases 
to  agricultural 
of  secondary  micro-  and  nanoplastics  from 
abrasion  or  maintenance  of  outdoor  plastic 
goods  and  coated  surfaces”,  say  Alimi  et  al. 
(2018),  and  may  subsequently  “undergo  various 
transformations 
commonly  associated  with 
natural  or  anthropogenic  colloids;  namely, 
interactions 
homo-  and  heteroaggregation, 
with  microorganisms  and  macromolecules  (e.g., 
adsorption  of  proteins,  natural  organic  matter) 
and biodegradation”.

enter  the  groundwater,  being  carried  down  by 
rainfall,  wet-dry  cycles,  earthworms  and  other 
soil  organisms  (Boyle  and  Örmeci,  2021;  Blasing 
and  Amelung,  2018).  Nanoplastic  particles  also 
have the potential to pass through soil pores and 
enter  groundwater  aquifers  (Rillig  et  al.,  2017; 
Chae and An, 2018). As PS nanoplastics have been 
shown to act as carriers for chemical pollutants – 
pyrene and ‘BDE 47’ – present in saturated soil, it 
is  possible  that  these  NPs  could  transport  these 
toxic  pollutants  into  the  groundwater  of  aquifers 
(Banerjee and Shelver, 2021).

Earthworms suffer detrimental physiological effects 
due  to  PE  microplastic  exposure  in  soil  including 
damaged immune systems and inhibition of sperm 
formation  in  male  worms  (Kwak  and  An,  2021), 
mortality,  reduced  burrows  and  reduced  growth 

48

(Huerta Lwanga et al., 2016), and gut inflammation 
(Rodriguez-Seijo et al., 2017). Studies examining the 
impact of nanoplastics on earthworms, meanwhile, 
are  scarce.  Kwak  and  An  (2021)  found  that  they 
excrete nanoplastic PE particles into the soil, after 
ingesting/digesting  microplastics.  This  breakdown 
of  microplastics  to  nanoplastics  has  been  shown 
to occur due to earthworm gut bacteria (Ng et al., 
2018)  –  a  phenomenon  also  demonstrated  in  krill 
(Dawson et al., 2018).

Having  established  that  earthworms  can  ingest 
microplastics, and excrete fragmented nanoplastic 
particles, it is likely that nanoplastics are also being 
moved further down into the soil depths via similar 
processes to microplastics. For example, movement 
of nanoplastic PS particles, of 100 nm size, has been 
observed  in  a  range  of  soil  types,  and  one  study 
found  that  the  higher  the  pH,  and  lower  the  iron/
aluminium oxide content, the greater the movement 
of  PS  nanoplastics  in  the  soil  (Boyle  and  Örmeci, 
2021).  Different  physico-chemical  compositions  of 
soil  and  the  array  of  different  nanoplastic  particle 
types  are  likely  to  interact  in  a  range  of  manners 
together – an area which requires further research. 

3.6.2 Terrestrial plants and fungi

Both  nanoplastics  and  small  microplastics  can  be 
taken up by terrestrial plants (Mateos- Cárdenas et al., 
2021; Larue et al., 2021) – raising concerns regarding 
foodchain  contamination.  Studies  comparing  uptake 
and accumulation of microplastics versus nanoplastics 
found that nanoplastics accumulate more in root tips 
of plants. Nanoplastics can be absorbed into the roots 
of  cress  (Lepidium  sativum),  broad  bean  (Vicia  faba) 
and onion (Allium cepa), for example, and translocate 
to the shoots of wheat (Triticum aestivum). Uptake and 
translocation of styrene maleic anhydride nanoplastics 
has  also  been  demonstrated  in  orange  jasmine 
(Murraya exotica) (Larue et al., 2021). A recent study 
investigated whether three tree species could take up 
nanoplastics, finding that nanopolystyrene was taken 
up by roots and into tissues (Murazzi et al., 2022).

Trophic chain transfer of fluorescent PS nanoplastics 
has been shown to occur from the soil into mung beans, 
and into African giant snails that feed on these. Studies 
of commercialised vegetables, grown in greenhouses, 
show  that  vegetables  and  cereals  also  accumulate 
high amounts of phthalates – additives in plastics – 
illustrating a high risk of transfer to consumers (Larue 
et al., 2021).

Nanoplastics  can  trigger  oxidative  stress  reactions 
in  plants,  affecting  germination,  photosynthesis  and 
growth. Additives leaching from nanoplastics are toxic 
to the genes and cells of some plants, (for example, 
Allium  cepa,  when  exposed  to  the  leachate  of  a 
biodegradable nanocomposite of PLA and nanoclays) 
(Larue  et  al.,  2021).  Nanoplastics  can  also  indirectly 

affect  plant  growth  by  changing  soil  properties  and 
the microbial community (Chen et al., 2022). 

Interestingly,  nanoplastics  in  soil  may  alleviate 
the toxic effects of other soil pollutants – such as 
Arsenic  (As)  and  Cadmium  (Cd)  –  on  crops,  which, 
when combined with PS nanoplastics were less toxic 
than when wheat was exposed to them alone. This 
was due to the nanoplastics attaching the As and 
Cd  to  their  surface,  causing  the  plants  to  be  less 
exposed to the toxins. Also, the nanoplastics reduce 
root activity, which could decrease the uptake of As 
and Cd (Larue et al., 2021).

Smaller  sized  and  positively  charged  particles 
are  potentially  more  toxic  to  plants.  For  example, 
one  study  found  positively  charged  nanoplastic 
particles  more  toxic  to  Arabidopsis  thaliana  than 
negatively charged nanoplastics (Sun et al., 2021). 
There are a number of other variables that require 
more  investigation,  especially  to  reflect  real-world 
particles and environments in findings. For example, 
aged particles of plastic have been shown to exert 
more toxic effects on phytoplankton (Box 12), so lab-
based studies using non-aged particles might show 
less toxic impacts. However, spherical shapes appear 
to be more easily absorbed by roots than particles 
of other shapes, while irregular, aged particles may 
be less bioavailable (Maity et al., 2022). Meanwhile, 
strong binding of nanoplastic to soil particles may 
also slow down uptake, but nanoplastics present in 
water in soil pores appear more readily available to 
plants (ibid.).

49

Biodegradation of nanoplastics in the rhizosphere – 
the region of soil around a plant’s roots – is yet to 
receive  much  research  attention.  One  study  found 
that  chemicals  exuded  from  roots  in  response  to 
nanoplastics  may  drive  their  degradation  in  soil 
(Yoon et al. 2021).

Nanoplastics  have  been  found  to  be  absorbed 
into  fungi  cell  walls,  showing  their  ability  to  cross 
barriers  considered  impermeable  for  many  other 
toxic  substances  (de  Souza  Machado  et  al.,  2018; 
EASAC,  2020).  Some  studies  have  indicated  that 

nanoplastics inhibit fungal enzyme activity, damage 
hyphae and affect fungal community structure (Du 
et  al.,  2022).  This  could  have  knock-on  effects  in 
ecosystems,  for  example  leading  to  lower  levels 
of leaf litter decomposition and nutrient release in 
freshwater streams. A recent review looking at the 
effects  of  micro-  and  nanoplastics  on  microbiota 
in  soil  and  water,  also  highlighted  their  complex 
effects  on  ecosystems,  and  individual  organism 
health (Santos et al., 2022).

3.6.3 Other terrestrial organisms

Contamination  of 
terrestrial  organisms  with 
microplastics  is  already  widespread,  with  Zhao 
et  al.  (2016),  for  example,  finding  microplastics  in 
the digestive tract of 94% of dead terrestrial birds 
in  China,  with  diverse  foraging  behaviour.  Smaller 
particles of plastic, microplastics and nanoplastics 
can  be  ingested  or  inhaled,  blocking  the  digestive 
tract,  or  abrading  and  irritating  mucosa  (EASAC, 
2020).  In  one  study,  nanoplastics  inhaled  by  rats 
were transported from the lungs to the capillaries, 

from where the particles could be distributed to the 
rest  of  the  body  (de  Souza  Machado  et  al.,  2018). 
Nanoplastics have now been shown to cross highly 
selective  membranes,  such  as  the  blood-brain 
barrier,  and  the  human  placenta  (Sökmen  et  al., 
2020;  de  Souza  Machado  et  al.,  2018;  Campanale 
et al., 2020).

50

4. Ecotoxicity of nanoplastics

Studies  have  demonstrated  a  number  of 
impacts  of  nanoplastics  on  terrestrial,  marine 
and  freshwater  organisms.  Toxicity  seems  to  be 
chiefly  due  to  damage  to  cell  membranes  and 
oxidative  stress  (increased  free  radicals  which 
cause damage to tissues). Findings in this field are 
mostly  laboratory-based,  using  model  organisms 
and  exposure  to  relatively  high  concentrations  of 
nanoplastic. At present, technical limitations make 

quantification  of  concentrations  of  environmental 
levels  of  nanoplastics  challenging,  but  innovative 
techniques  are  emerging  that  permit  detection 
of  nanoplastics  in  field  samples.  Establishing 
evidenced-based  environmental  concentrations 
of  nanoplastics  will  further  validate  lab-based 
mechanistic  studies  of  ecotoxicity,  coupled  with 
determinations of accurate concentrations in both 
media and tissue samples.

Box 13: Not all nanoplastics are (equally) toxic

Some  plastics  –  and  by  extension  micro-  and 
nanoplastic  particles  –  are  more  hazardous  to 
health  than  others.  While  the  surface  properties 
and size of nanoplastics may affect their toxicity, 
and  risk  is  related  to  extent  of  exposure,  their 
chemical constituents are also key.

Constituent monomers

looked  at  chemicals  used 

Researchers 
in  55 
thermoplastic and thermosetting polymers to develop 
a  hazard  ranking  of  plastic  polymers  for  the  EU 
classification and labelling regulation (CLP) (Lithner et 
al., 2011). According to the ranking, the most hazardous 
–  polyurethanes,  polyacrylonitriles,  PVC,  epoxy  and 
styrenic  copolymers  –  are  made  of  monomers 
classified as mutagenic and/or carcinogenic.

Additives

Some  plastics  are  made  of  monomers  considered 
non-hazardous (e.g. polyethylene and polypropylene), 
but  which  contain  harmful  additives,  for  instance 
triclosan, phthalates, bisphenol A and formaldehyde. 
Leachates from plasticised PVC and epoxy products 
have been shown to be the most toxic to Daphnia 
magna  (Lithner  et  al.,  2009;  Lithner  et  al.,  2012), 
chiefly  due  to  hydrophobic  organic  compounds. 
A  recent  study  of  eight  polymers  found  that  all 
chemicals  leached  into  water  were  toxic  to  some 
degree, with LDPE, PVC and polyurethane inducing 
most toxic effects (Zimmerman et al., 2021). 

Plastic additives include endocrine disruptors on the 
Priority List22. A recent Expert Consensus Statement 
outlined  their  key  characteristics  as  a  basis  for 
hazard identification (La Merrill et al., 2020), but no 
hazard ranking has yet been compiled. Under the EC 
Chemical Sustainability Strategy, the EU will create 
a legal definition for endocrine disruptors, splitting 
them  into  two  categories  –  known/presumed  and 
suspected,  further  subdivided  into  human  health 
and  environmental  endocrine  disruptors  –  so  that 
they  may  be  addressed  in  CLP  (classification, 
labelling and packaging) regulation. 

Attached contaminants

Persistent organic pollutants (e.g. PCBs and PBDEs) 
and  heavy  metals  can  attach  to  nanoplastic 
particles  (Davranche  et  al.,  2019),  making  them 
more toxic to species that ingest them. For instance, 
PVC  with  triclosan  caused  mortality  in  lugworms 
(Arenicola  marina)  in  one  study  (Browne  et  al., 
2013);  in  another,  the  toxicity  of  polyethylene  to 
fish  increased  when  combined  with  contaminants 
including metals (Rochman et al., 2013). The degree 
to which such mixing occurs on the surface of micro- 
and nanoplastics, under different conditions, is still 
being studied (Yu et al., 2019). 

22  Candidate List of substances of very high concern for Authorisation - ECHA (europa.eu)  

Available at: https://www.echa.europa.eu/candidate-list-table

51

Rate of leaching

Knowledge gaps

Physical conditions can enhance the release of toxic 
additives  from  plastics.  For  instance,  researchers 
found that UV radiation enhanced toxicity of low-
density  polyethylene  recyclate  (LDPE-R),  starch 
blend (SB), bio-based polybutylene succinate (Bio-
PBS)  and  PVC,  but  had  little  or  no  effect  on  PET, 
polystyrene, PP and LDPE (Klein et al., 2021).

Numerous  studies  reveal  toxic  effects  of  plastic 
is  no  comprehensive 
leachates.  However,  there 
characterisation  of  the  complex  chemical  mixtures 
present  in  plastics;  consumer  plastics  contain  toxic 
compounds that remain unidentified and unregulated 
(Zimmerman  et  al.,  2019;  Gunaalan  et  al.,  2020). 
Plastic  and  nanoplastic  toxicity  studies  also  tend  to 
focus  on  organisms  such  as  plankton  (e.g.  Daphnia), 
zebrafish  (Danio  rerio)  and  earthworms  (Annelida), 
whose metabolism may not reflect that of mammals 
(Shen et al., 2019).

Our study demonstrates that consumer plastics contain compounds 
that are toxic in vitro but remain largely unidentified. Since the risk 
of unknown compounds cannot be assessed, this poses a challenge to 
manufacturers, public health authorities, and researchers alike.  
However, we also demonstrate that products not inducing toxicity  
are already on the market 

Zimmerman et al., 2019

4.1 Potential mechanisms of toxicity 

The  toxicity  of  nanoplastic  particles  (  at  certain 
levels  of  exposure)  is  suggested  to  be  due  to  a 
number  of  factors:  membrane  damage,  oxidative 
stress, 
immune  response  and  genotoxicity  – 
with  cytotoxicity  (cell  destruction)  being  mostly 
due  to  membrane  damage  and  oxidative  stress 
(Bhattacharjee  et  al.,  2014).  In  one  experiment, 
polyethylene nanoparticles penetrated the plasma 
membrane  bilayer,  causing  structural  changes  in 
these biological membranes (Banerjee and Shelver, 
2021).  Another  study  showed  that  particles 
of  nanoplastics  that  entered  into  cells  caused 
internal  membranes  to  become  more  permeable, 
and  interacted  with  cellular  organelles  (Yong  et 
al.,  2020).  Reactive  oxygen  species  (ROS)  can  be 
generated  within  cells  upon  processing  of  the 
plastic  particles,  causing  cellular  stress  (Rubio  et 
al.,  2020).  Smaller  nanoplastic  particles  can  also 
cross the gut/lung barrier – triggering intracellular 
oxidative  stress,  followed  by  cytotoxicity  in  the 
organs  in  which  the  nanoplastics  accumulate. 

Both  particle  or  ROS  translocation 
into  the 
nucleus of a cell can damage DNA replication and 
repair  machinery,  contributing  towards  particle 
genotoxicity (ibid.).

In mammalian cells nanoplastics can cause nuclear 
membrane  disruption,  oxidative  stress,  release 
of  damage  associated  molecular  patterns  and 
cascading  effects  of  inflammation  and  cell  death 
pathways. Liver cells of mice have been shown to 
suffer ROS, due to presence of 50 nm PS particles, 
as  well  as  cell  damage.  Human  liver  cancer  cells 
had  a  similar  reaction  with  generation  of  free 
radicals, and mouse macrophages – immune cells – 
destroyed themselves in the presence of positively 
charge  PS  beads.  Most  studies  suggest  oxidative 
stress  is  triggered  by  nanoplastics,  however,  one 
type of toxicity mechanism may trigger a cascade 
of  other  toxicity  mechanisms  as  they  are  all 
interconnected (ibid.).

52

At present a thorough understanding of the effect 
of  nanoplastics  on  humans  and  other  organisms 
is  incomplete  –  comprehensive  studies  using 
diverse plastic materials at relevant environmental 
concentrations, modelling for chronic exposure, are 
needed for a realistic hazard and risk assessment 
(Lehner  et  al.,  2019).  In  addition,  the  terrestrial 
environment, and freshwater ecosystems are less 
well  studied  than  marine  habitats,  and  require 
more  extensive  research  focus  in  the  future. 
terrestrial,  and  freshwater 
Studies 
ecosystems  will  aid  our  overall  understanding  of 
nanoplastic  pollution  in  the  environment,  and  its 
possible impact on human health.

involving 

Nanoplastic toxicity is likely initiated by a cascade 
of  changes  at  the  subcellular  level  –  such  as 
oxidative  stress  –  that  then  propagate  through 
the  biological  hierarchy,  impacting  for  example 
migratory  behaviour,  reproduction  success  and 
foraging  behaviour,  or  causing  death  (de  Souza 
Machado  et  al.,  2018).  Effects  and  the  extent 
of  toxicity  seems  to  depend  on  a  wide  range  of 
variables,  including  polymer  type,  presence  of 
additives,  size,  shape,  surface  charge  of  plastics, 
and dose (Yee et al., 2021; Larue et al., 2021).

Genotoxic and mutagenic effects – changes to DNA 
– have been demonstrated, with organisms exposed 
to polystyrene nanoparticles exhibiting altered gene 
expression,  for  example,  chromosome  ruptures 
and nuclear abnormalities (Guimaraes et al., 2022; 
Alaraby et al., 2022; Brandts et al., 2022). A study 
looking at the effects of polystyrene nanoparticles 
on  human  white  blood  cells  also  found  DNA 
damage,  in  addition  to  changes  in  the  expression 
of cytokines related to inflammation, immune and 
stress responses (Ballesteros et al., 2020). 

Immunotoxicity  –  effects  on  cells  related  to  the 
immune  system  –  has  been  shown  in  a  number 
of  experiments.  Sea  urchins 
(Paracentrotus 
lividus)  exposed  to  polystyrene  nanoparticles 
exhibited  reduced  phagocytosis  (destruction  of 
foreign  particles),  with  a  more  significant  effect 
from  negatively  charged  particles  (Murano  et  al., 
2021;  see  4.5);  another  study  suggested  that 
polystyrene  and  polycarbonate  nanoparticles  act 
as  stressors  on  the  immune  system  of  fathead 
minnows (Greven et al., 2016). Brandts et al. (2018) 
found  that  nanoplastics  affect  the  liver  and  lipid 
metabolism of fish, interfering with their ability to 
use  energy  reserves,  and  also  affect  skin  mucus 
which functions as a barrier against pathogens.

Meanwhile,  some  researchers  highlight  the  way 
nanoplastics  may  act  as  a  ‘Trojan  Horse’  for 
other  potentially  toxic  substances,  for  example  
toxictrace  metals  attached  to  their  surface(e.g. 
Hildebrandt  et  al.,  2021;  Domenech  et  al,  2021). 
Although  secondary  nanoplastics  are  different 
from 
intentionally  manufactured  nano-sized 
polymers,  it  is  worth  noting  that  there  has  been 
research  into  the  potential  for  using  the  latter 
to  deliver  drugs  to  intercellular  compartments, 
highlighting their ability to act as carriers of other 
chemical substances. 

Studies  demonstrating  nanoplastic  particles 
crossing  cell  membranes  and  translocating  –  in 
fish,  invertebrates,  plants,  fungi  and  humans  – 
are beginning to be published (e.g. Sökmen et al., 
2020; de Souza Machado et al., 2018; Campanale 
et  al.,  2020)  (see  section  3.5  and  3.6)  The  exact 
translocation  mechanisms  (for  example,  between 
cells  of  the  gut  wall  epithelium,  or  through  cells) 
remain unclear. 

… the question arises whether nanoplastics are able to reach deeper 
organs besides the intestine due to an altered cellular fate 

Paul et al. 2022

 
53

4.2 Effect concentrations

To  understand  whether 
concentrations  of 
nanoplastics  in  the  environment  pose  a  risk  of 
toxic effects, knowledge on the levels at which they 
negatively  affect  organisms  is  needed.  A  growing 
number of laboratory studies are being published, 
offering data on the effects of exposure to different 
‘doses’  of  nano-  and  microplastics.  Besseling  et 
al.  (2019)  reviewed  data  on  exposure  via  water 
(mg/ml)  and  sediment  (g/kg/dry  weight)  across 
brackish, marine and freshwater ecosystems. Effect 
concentrations  reported  in  this  review  are  shown 
in  Table  5,  which  compiles  findings  for  different 

plastics. The levels vary widely – perhaps because 
different  types  of  polymer  have  very  different 
ecotoxic effects, depending on size, particle charge, 
and presence of chemicals such as plasticisers.

The  review  highlights  the  limited  data  available 
on  effect  concentrations  of  nanoplastics 
in 
sediments; most studies to date have focussed on 
concentrations  of  nanoplastics  in  water.  There  is 
also a knowledge gap on the potential impacts of 
nanoplastics on the environment, when contained 
in food or sediment. 

Table 5. Summary of published effect data of nanoplastic on organisms. LC50: lethal dose 50%; EC50: effect 
concentration  50%;  LOEC:  lowest  observed  effect  concentration;  NOEC:  no  observed  effect  concentration 
(Redrawn from Source: Besseling et al., 2019).

Exposure 
Medium

Ecosystem

LC50 – level 
at which 
50% of 
organisms 
die

EC50 – level 
at which 
50% of 
organisms 
are affected

LOEC – 
lowest level 
at which 
effects are 
noted

NOEC – 
no effects 
observed 
at this 
level

Water 
(mg/L)

freshwater

4 - 36

0.5-1.6

4.5-1 x 103

0.5 - 1

brackish

0.2 - 22

marine

0.8 - 3.9

Sediment 
(g/kg DW)

freshwater

brackish

marine

-

-

-

-

13

-

-

-

-

1 - 313

0.1 - 250

10 - 100

1

-

-

-

-

-

 
54

4.3 Species sensitivity distributions and preliminary threshold concentrations 
for nanoplastics

Toxic  effects  of  chemicals  are  often  reported  for 
tests with single species and such tests inform the 
determination of predicted no-effect concentrations 
(PNEC),  taking  account  also  of  any  data  available 
on  other  species  and  the  level  of  uncertainty 
regarding their sensitivity. At concentrations below 
the PNEC, no adverse effects on the ecosystem are 
expected.  Therefore,  PNEC  values  often  form  the 
basis for setting protective thresholds such as the 
Environmental  Quality  Standards  under  the  Water 
Framework Directive. 

in 

If  test  results  for  several  taxonomic  groups  are 
available, the results from single species tests can 
be  combined  in  species  sensitivity  distributions 
(SSDs).  In  Figure  18  below,  Besseling  et  al.  (2019) 
have  provided  information  on  SSDs  for  organisms 
affected  by  nanoplastics 
the  water.  For 
nanoplastics, the most sensitive species was found 
to  be  the  copepod,  Tigriopus  japonicus,  (50  nm 
spherical  PS  particles)  and  the  least  sensitive  the 
algae  Scenedesmus  obliquus  (70  nm  spherical  PS 
particles). The study finds a hazardous concentration 
(or HC5, where 5 per cent of species in the SSD exhibit 
an  effect)  for  nanoplastic  of  5.4  mg/L  –  identical 
to that of microplastic. Using an assessment factor 
calculation,  the  researchers  proposed  a  PNEC  of 
1.1mg nanoplastic/L water. 

Fig.  18.  Species  sensitivity  distributions  of  organisms  from  marine,  estuarine  and  freshwater  environments 
exposed  to  microplastic  (Panel  A)  or  nanoplastic  (Panel  B)  via  the  water  phase.  Effect  thresholds  represent 
chronic LOECs. Grey curves represent the 95% confidence intervals. Source: Besseling et al. (2019)

 
55

When  contaminants  such  as  metals,  antibiotics, 
pathogenic  micro-organisms  (e.g.  Vibrio  spp)  and 
persistent  organic  pollutants  (POPs)  attach  to 
nanoplastic  particles,  those  contaminants  may 
become less toxic in the wider environment, through 
being  less  bioavailable  (ibid;  Banerjee  and  Shelver 
et  al.,  2021).  In  some  cases,  however,  nanoplastic 
may  enhance  the  toxicity  of  other  contaminants, 
for  example  if  particles  are  positively  charged 
(Shen et al., 2019), though the reason for increased 
toxicity in some experiments is not always explained 
(Campanale et al., 2020). Indeed, there is a lack of 
knowledge on nanoplastic corona formation, yet the 
surface structure of nanoplastics in the environment 
must  be  taken  into  account  when  attempting  to 
predict  their  interactions  with  cell  membranes 
–  a  complex  challenge  (Kihara  et  al.,  2021). 

4.4 Eco-corona

Nanoplastics can adsorb (attach) both environmental 
pollutants  and  naturally  occurring  substances  (e.g. 
natural organic matter or extracellular proteins) onto 
their surface. This adsorbed material is often referred 
to  as  a  corona  since  natural  organic  matter  (NOM) 
or extracellular proteins may form a layer (a corona) 
covering the particle surface. The term ‘eco-corona’ 
is  used  to  refer  to  an  ecological  molecule  corona 
comprised of substances from the environment. 

As  with  other  aspects  of  nanoplastic  behaviour, 
characteristics  of  the  medium 
influence  eco-
corona  formation  (ions,  salinity,  organic  matter, 
pH), and the eco-corona in turn influences the way 
nanoplastic particles interact with organisms in that 
environment.  For  example,  when  humic  acid  was 
added to a medium containing PS nanoplastics, the 
nanoplastic  became  less  toxic  to  algae  because  it 
did not adhere to the algae as much as in a control 
medium (Oberbeckmann and Labrenz, 2020). 

[T]he particle’s “biological identity” should consider the full complexity  
of the surface structure 

4.5 Electrostatic charge

Source: Kihara et al., 2021

Fig. 19. Potential toxicity and uptake mechanisms of nanoplastics in animal and fungal cells. Carboxyl and amino 
terminated or lecithin coated polystyrene nanoparticles have diverse cellular fates, that influence the toxicity 
mechanism.  Showing  surface  properties  of  nanoparticles  are  important  in  determining  their  level  of  toxicity. 
Mechanisms so far reported do not fully explain toxicity and uptake, further mechanisms need to be identified. 
Source: de Souza Machado et al. (2018)

 
 
56

Nanoplastic  particles  can  attach  externally  to 
cell  membranes  and  cause  toxicity  by  disrupting 
essential membrane processes of an organism (de 
Souza Machado et al., 2018). The surface chemistry 
and electrostatic charge of nanoplastics determines 
their loading capacity for chemicals, and ability to 
cross  cell  membranes,  influencing  toxicity  (ibid; 
Roach et al., 2006). For example, one experiment 
showed that positively charged polystyrene beads 
of  50  nm  –  because  of  the  positive  amino  end 
molecule  branch  (NH2)  –  can  stick  to  cells  and 
cause  high  toxicity  to  yeast  at  concentrations  of 
around 10 mg/L (Miyazaki et al., 2014). Conversely, 
negatively charged PS nanoplastics – with negative 
carboxyl end molecules (COOH) – had no effect on 
the growth rate of cells in this experiment.

The  acute  toxicity  of  positively  charged  PS 
nanoplastics is due to their electrostatic attraction 
to negatively charged cell walls. Positively charged 
particles  of  amino  modified  PS  nanoplastics  –  of 
60  nm  or  400  nm  –  also  triggered  greater  lung 
responses,  and  thrombosis, 
in  hamsters  and 
rabbits  (Hamoir  et  al.,  2003;  de  Souza  Machado 

et  al.,  2018).  However,  it  has  been  shown  that 
the  surface  charge  of  nanoplastics  particles  can 
be  neutralised  by  the  occurrence  of  eco-coronas 
(see  section  4.4.).  In  their  2017  review,  Pulido-
Reyes  et  al.  found  that  eco-coronas  reduce  the 
aggregation  potential  and  bio-reactivity  of  some 
nanoparticles.  Natarajan  et  al.  (2020)  clearly 
indicated  a  significant  decline  in  toxic  effects 
of  PS  nanoplastics  on  marine  microalgae  due  to 
formation  of  an  eco-corona,  and  Saavedra,  Stoll 
and  Slaveykova  (2019)  demonstrated  that  eco-
coronas modulated the surface charge and reduced 
toxicity  to  zooplankton.  This  highlights  the  need 
to  examine  nanoplastic  interactions  in  a  natural 
environment,  as  well  as  in  laboratory  conditions, 
for  a  full  understanding  of  their  effects.  The 
chemical  reactivity  and  electrostatic  interactions 
of  nanoplastics  are  not  the  only  factors  involved 
in toxicity; nanosized particles of inert substances 
such as gold can translocate to organs and trigger 
negative effects in biological systems (Prüst, Meijer 
and Westerink, 2020). 

4.6 Particle and chemical toxicity in plants

Leachates  from  plastic  degradation  may  contain 
polymer  monomers  and/or  additives  that  can 
cause toxicity to plants. For example, genotoxicity 
and  cytotoxicity  were  induced  in  onions,  Allium 
cepa,  exposed  to  leachates  from  the  composting 
of a new biodegradable nanocomposite of PLA and 
nanoclays (Larue et al., 2021).

PS  nanoparticles  have  been  shown  to  have  an 
effect  on  the  cell  division  (genetic  functioning)  of 
seedlings of two species of plants, with nanoplastics 
detected inside the cells of one of the species, A. 
cepa,  affecting  organelles  of  these  seedlings.  An 
impact  on  lipid  mobilisation  in  exposed  seedlings 
was also detected due to PS nanoplastic exposure. 

Oxidative  stress  was  demonstrated  in  A.  cepa 
PS  nanoplastic-exposed  seedlings;  however,  the 
seedlings  appeared  to  be  able  to  prevent  this 
being harmful – up to a certain dose. The authors 
suggest  this  implies  the  mechanism  for  cell  and 
gene  toxicity  in  A.  cepa  exposed  to  PS  beads,  is 
not  caused  by  the  oxidative  stress,  but  rather  a 
different mechanism (Larue et al., 2021).

In  wheat  plants,  PS  nanoparticles  affected  the 
metabolic  system  of  these  plants  –accelerating 
carbohydrate  and  amino  acid  metabolism  –  a 
mechanism used to cope with PS toxicity (Larue et 
al., 2021). PS nanoplastics caused reactive oxygen 
species  formation,  and  influenced  genes  related 
to  oxidative  stress,  when  Arabidopsis  thaliana, 
(mustard  family,  includes  cabbage  and  radish) 
was exposed to them. In addition, changes in root 
morphology,  nutrient  imbalances,  and  alterations 
in  plant  root  to  shoot  morphology,  have  been 
reported in the literature (ibid.).

Micro-  and  nanoplastics  may  also 
carry 
contaminants that could cause toxicity in plants and 
other organisms – for example, polycyclic aromatic 
hydrocarbons  (PAHs),  polychlorinated  biphenyls 
(PCBs),  perfluoroalkyl  acids  (PFAAs),  persistent 
organic  pollutants  (POPs),  pharmaceuticals  and 
(PPCPs),  and  metals 
personal  care  products 
(Hartmann et al., 2017).

4.7 Chemical additives in nanoplastics

Several  studies  have  reported  that  microplastics  – 
and likely nanoplastics – can also cause toxic effects 
due  to  chemicals  present  in  the  plastic  (Chae  and 
An,  2018).  Plastics  can  contain  additive  chemicals 
such as plasticisers (BPA, DHEP, DiNP), antioxidants, 
UV  stabilizers  and  flame  retardants,  which  are 
then  released  into  aquatic  environments,  causing 
harm  to  aquatic  organisms.  This  plastic  leachate 
has  been  demonstrated  for  both  microplastics  and 
nanoplastics  (Zhao  et  al.,  2017;  Larue  et  al.,  2021). 
These contaminants may be transferred through the 
food chain to predators at upper trophic levels and, 
as such, their presence could pose a threat to aquatic 
organisms and ecosystems, and human health. 

(and 

levels 

Plastic  additives  have  been  found  at  moderately 
likely 
in  microplastic-rich 
high 
nanoplastic-rich)  sludge  from  water  treatment 
plants,  used  for  agricultural  purposes  (EASAC, 
2020;  Boyle  and  Örmeci,  2021),  with  potential 
to  impact  plant  health.  Researchers  have  not  yet 
quantified  the  release  of  additives  into  soil  or 
water (Larue et al., 2021). 

57

Capolupo et al. (2020) investigated the composition, 
and effect, of leachates from car tyre rubber (CTR), 
PP,  PET,  PS,  and  PVC.  These  leachates  –  which 
included  plasticisers,  antioxidants,  antimicrobials, 
lubricants,  and  vulcanisers  –  were  derived  from 
microplastic  particles  rather  than  nanoplastic 
particles;  however,  the  impact  of  the  leachates 
themselves  is  likely  unaffected  by  the  particle 
size. The leachates negatively affected membrane 
stability,  fertilisation  and  embryo  development  in 
Mediterranean mussels (Mytilus galloprovincialis) – 
with 50% of the embryos dying at concentrations 
ranging  from  0.7%  (CTR)  to  65%  (PET)  of  the 
total  leachate.  Notable  additives  in  the  different 
types  of  polymer  included:  benzothiazole  (CTR), 
phthalide  (PVC),  acetophenone  (PP),  cobalt  (CTR, 
PET), zinc (CTR, PVC), lead (PP) and antimony (PET). 
This study highlights that chemical composition of 
plastic leachates is implicated in toxicity to marine 
biota – with increasing contamination by additives 
causing increasing toxicity.

58

5. Impacts on human health

Constant  human  interaction  with  plastic  items 
leads  to  mouth  and  skin  exposure,  as  well  as 
inhalation.  Assessing  the  risk  to  human  health 
from  nanoplastics  specifically,  is  a  complex  task. 
For example, different plastic materials may have 
different effects on humans and other organisms in 
the food chain; and potential sources of exposure 
(food,  personal  care  products,  city  dust,  clothing) 
may  pose  different  risks.  However,  some  general 
patterns  of  impact  of  nanoplastics  (and  smaller 
microplastics)  can  be  noted.  Firstly,  the  potential 
hazard  of  nanoplastic  particles  increases  with 
decreasing size, due to increased bioavailability in 
the body, and potential to pass through biological 

5.1 Routes into the human body

membranes (Paul et. al., 2022). The type of hazard 
posed by larger particles is likely to be mechanical 
damage,  followed  by 
inflammation,  whereas 
smaller  particles  which  could  translocate  to  body 
organs,  may  accumulate  causing  oxidative  stress 
and  inflammation  more  systemically.  In  addition, 
quantification  and  detection 
for 
nanoplastics  still  need  to  be  refined.  Considering 
this, it is understandable that there is no consensus 
on which aspects of nanoplastics are most harmful 
to humans and other organisms, however, research 
is underway to address this knowledge gap. 

techniques 

Fig. 20. Key routes by which plastic particles enter the human body. Source: Yee et al. (2021)

It is estimated that, on average, an adult person consumes around 
39,000–52,000 particles a year or 5 g of plastic every week — the 
equivalent of a credit card 

59

Schwarzfischer and Rogler, 2022

translocation and potential health risks in humans. 
It is important to recognise that it is not possible 
to extrapolate directly from studies on non-human 
species, however current assertions of impacts are 
theoretical  assumptions,  based  on  animal  model 
studies and lab-based data (Yee et al., 2021; Prüst, 
Meijer and Westerink, 2020). Meanwhile, detecting 
micro- and nanoplastics in human tissues presents 
its  own  issues.  For  example,  a  study  that  found 
microplastics 
in  human  placenta  and  foetal 
meconium  highlighted  that  contamination  of 
samples – for example from airborne fallout – is a 
methodological challenge (Braun et al, 2021).

Microplastics have been identified in human stool, 
but risks from ingestion are not yet fully understood 
(Schwabl et al., 2019; Zhang et al., 2021). Ingestion 
of nanoparticles and nanoplastics, is an emerging 
and complex research field, where the size, shape, 
charge,  and  surface  modification,  as  well  as  the 
human membrane cell type encountered influence 
the mode of cellular uptake and entry into the body 
(Foroozandeh  and  Aziz,  2022).  Nanoplastics  may 
interact with proteins, lipids, carbohydrates, nucleic 
acids,  ions  and  water  in  the  human  body  leading 
to the formation of coronated nanoplastics which 
increases  their  absorption  and  translocation  (Yee 
et  al.,  2021).  It  is  likely  that  nanoplastic  particles 
and  very  small  microplastic  particles  have  the 
potential  to  be  ingested  in  the  human  gut,  due 
to  their  material  characteristics  conferring  higher 
bioavailability, for example, particles below 2.5 µm 
can be absorbed by M-cells or Payer’s patches in 
the human gut (Schwarzfischer and Rogler, 2022). 
Furthermore, tight junction pores between cells in 
the human gut, have a functional size of 1.5 nm – 
enabling nanoplastics to potentially enter via this 
route – whereas microplastics would be too large.

Studies  examining  the  ingestion  of  nano  plastic 
particles  in  the  gut  have  noted  in  models  of 
human digestion, that ingestion and translocation/
accumulation  of  NP  and  MP  occurs  in  organs 
distant  from  the  gut  (Schwarzfischer  and  Rogler, 
2022). It has been suggested that those suffering 

Fig.  21.  Nano-  and  microplastics  contamination 
of  fresh  and  salt  water.  Secondary  microplastic 
particles  emerge  with  plastic  fragmentation  whilst 
in  the  environment.  UV  radiation,  mechanical  forces, 
biological  degradation,  and  embrittlement  result  in 
the  formation  of  nanoplastic  particles,  which  are 
ingested  by  marine  zooplankton  and  consequently 
enter the marine and human food chain. Nanoplastics 
may be consumed by people in contaminated drinking 
water,  and  other  foodstuffs.  However,  the  impacts  of 
ingested  plastic  particles  on  human  health  are  not 
yet  fully  understood.  Credit:  Marlene  Schwarzfischer, 
Department  of  Gastroenterology  and  Hepatology, 
University Hospital Zurich, University of Zurich, Zurich, 
Switzerland. Source: Schwarzfischer and Rogler (2022)

Nanoplastics may enter the human body through 
ingestion  of  contaminated  food  and  water, 
inhalation  of  airborne  particles  from  textiles  or 
polluted  outdoor  air,  or  penetration  of  particles 
through  wounds  or  weakened  skin  barriers  (Yee 
et  al.,  2021;  Schwarzfischer  and  Rogler,  2022). 
Research  has  indicated  nanoplastics  may  be 
consumed  via  contaminated  marine  animals, 
other  foods,  toothpaste,  beer,  honey,  salt,  sugar 
and  drinking  water  (ibid;  Koelmans  et  al.,  2019; 
Al-Sid-Cheikh et al., 2018). Medical implants such 
as  joint  replacements  may  also  be  a  source  of 
internal exposure.

of 

and 

nanoplastics 

subsequent 
Uptake 
translocation to organs was reported decades ago 
in rodents, and more recently in a number of other 
taxa, including in the brains of fish (Yee et al., 2021; 
Sökmen  et  al.,  2020).  However,  at  present  there 
is  a  paucity  of  systematic  studies  on  nanoplastic 

60

with conditions which effect the permeability of the 
gut  lining  –  such  as  inflammatory  bowel  disease 
or  crohns  disease  –  may  be  more  susceptible  to 
nanoplastic  particles  being  ingested  in  the  gut, 
due  to  disruption,  and  increased  permeability  of 
the  intestinal  epithelial  barrier.  At  present,  there 
is  insufficient  evidence  to  determine  whether 
nanoplastics  cause  inflammation  in  the  gut  of 
people suffering with inflammatory bowel disease, 
or crohns disease, due to conflicting findings in the 
research literature. 

The  WHO  recently  suggested  microplastics  (they 
did not include nano-sized plastics in their analysis) 
in  drinking  water  currently  have  no  discernible 
overt health risks, though they added a caveat that 
this  may  change  as  concentration  levels  increase 
– or as more studies are conducted that elucidate 
impacts (Mitrano and Wohlleben, 2020). At present 
in  vitro  studies  and  animal  models  have  shown 
harmful  effects  of  nanoplastics;  for  example, 
Lehner et al. (2019) found human cell lines showed 
evidence that nanoplastics are taken up and induce 
oxidative  stress  or  pro-inflammatory  response. 
However,  researchers  at  Arizona  State  University, 
identified  monomers  -  plastic  building  blocks  -  in 
all 47 human tissue samples examined,taken from 
Alzheimer’s patients. Toxic plasticizer, Bisphenol A 
(BPA),  still  used  in  some  food  containers  outside 
of  Europe  –  where  its  use  is  banned  –  was  also 
present  in  all  47  human  tissue  samples.  It  is 
unclear  how  these  monomers  of  plastic  came  to 
be  present  in  these  human  liver  and  fat  tissues, 
however, these two mammalian organs are likely to 
be exposed to plastic monomers (possibly resulting 
from micro or nanoplastic ingestion), as they filter 
or  collect  waste  products  from  the  human  body. 
The researchers suggested it to be concerning that 
these  non-biodegradable  plastic  materials,  that 
are present everywhere, can enter and accumulate 
in  human  tissues,  however,  at  present  the  health 
effects remain unknown (ACS, 2020).

in 
Microplastic  fibres  have  also  been  found 
human  lungs,  and  lung  cancer  tissue  (Jenner  et 
al.,  2022;  Pauly  et  al.,  1998),  and  studies  have 
indicated  that  microplastic  inhalation  may  cause 
inflammation  and  increase  risk  of  respiratory 
diseases  (Eschenbacher  et  al,  2003;  Mastrangelo 
et al., 2003). Prata, estimated that human beings 
inhale 26–130 MP particles per day, while studies 
by Vianello et al. suggested an average intake of 

272 MP particles a day (Prata, 2018; Vianello et.al, 
2019.  Atmospheric  microplastics  have  remained 
largely  unstudied  until  the  last  few  years  (and 
less  so),  so 
atmospheric  nanoplastics  even 
there  is  currently  only  limited  understanding  of 
atmospheric plastic pollution. 

In  outdoor  environments,  exposure  could  happen 
through breathing in contaminated aerosols from 
ocean waves or airborne fertiliser particles from dried 
wastewater treatment plants; indoor environments 
also contain airborne plastic particles, for example 
from synthetic textiles (Yee et al., 2021). A recent 
review on atmospheric microplastics (Zhang et al., 
2020)  concluded  that  the  effects  of  atmospheric 
microplastics, their chemical components and their 
adsorbed  pollutants  on  human  and  ecosystem 
health is unknown, but that the potential of micro- 
and  nanoplastic  to  influence  health  is  of  concern 
(Lehner et al., 2019; Wright and Kelly, 2017). 

to 

that  due 

review  asserts 

Another 
their 
hydrophobicity,  nanoplastic  should  be  repelled 
and  eliminated  in  mucus  (Rubio,  Marcos  and 
Hernández,  2019).  If  nanoplastics  were  able  to 
pass  through  the  physiological  tissue  barrier  in 
human  lungs  (which  separates  air  from  blood), 
they  could  be  translocated  to  other  parts  of 
the  body  –  however  the  risk  of  this  is  unknown. 
Neither  is  it  known  whether  nanoplastics  might 
pass  through  the  cerebral  endothelial  membrane 
(blood-brain  barrier).  It  is  thought  the  main  route 
of exposure that may present a risk to health is via 
the gastrointestinal tract, in particular via particles 
that have precipitated onto food from air ((Rubio, 
Marcos  and  Hernández,  2019.).  Indeed,  a  study 
that  found  microplastics  in  human  breastmilk, 
suggesting  it  had  translocated  into  blood  via  one 
of  the  routes  of  exposure,  found  no  significant 
relationship to variables including use of personal 
care  products  containing  plastic  compounds,  nor 
consumption of fish/shellfish, beverages, or food in 
plastic packaging (Ragusa et al., 2022). 

Further research is required to identify potential 
exposure, 
in  conjunction  with  standardised 
model  systems  to  evaluate  toxicity  and  long-
term  effects  in  humans  –  prerequisites  for 
human health risk assessments (Brachner et al., 
2022; Zarus et al (2021).

61

5.2 Toxicity in humans

Although toxicity evaluations have been conducted 
on  a  range  of  marine  species  and  rodents  (see 
section 3.5 and 4.1), to date there have been few 
studies  that  specifically  assess  the  impacts  of 
nanoplastic particles on human health, or how they 
move  through  the  gut,  lungs  and  skin  epithelia 
(Yee  et  al.,  2021).  Toxic  effects,  demonstrated  in 
experiments on organisms at lower trophic levels, 
cannot  be  directly  extrapolated  to  humans,  since 

the  metabolic  system  of  these  organisms  is 
quite  different,  and  the  ability  of  clearance  and 
resistance  to  nanoplastics  will  likely  differ.  The 
release  of  toxic  pollutants  from  nanoplastics  is 
also affected by many factors, and it is not certain 
whether such substances would be released in the 
human body in the same way as in other species 
(Shen et al., 2019). 

"Illustration of human body (©Shutterstock, photo by jijomathaidesigners)"

In addition, most studies to date have used large 
doses  of  nanoplastics  over  short  time  periods, 
rather  than  lower,  more  environmentally  relevant 
doses,  over  long  time  periods.  The  latter  is 
needed,  to  provide  a  more  accurate  analysis  of 
the  potential  impacts  of  nanoplastics  on  human 
health from current environmental contamination 
sources  (ECHA,  2020;  Shen  et  al.,  2020;  Yee  et 
al., 2021). However, as there is some evidence of 
bioaccumulation of nanoplastics in animal models, 
and  accumulation  of  plastic  monomers  in  human 
organs,  it  has  been  suggested  similarly  high 
concentrations  may  be  achieved  in  human  cells 
over time (Mukherjee, et. al., 2021; ACS, 2020). 

In vitro and in vivo studies have demonstrated that 
by  increasing  concentrations  of  nanoplastics,  the 
genotoxic,  inflammatory  and  cytotoxic  responses 
increase – similar concentrations may be reached 
with  the  bioaccumulation  process  (Mukherjee, 
et.  al.,  2021).  The  risk  to  human  health  requires 
investigation: studies on rats have shown inhaled 
nanoplastics may cause inflammation in the lungs 
(Lim et al., 2021); offspring of mice exposed to PS 
nanoparticles  during  pregnancy  exhibit  abnormal 
brain  development  (Jeong  et  al.,  2021);  and 
nanoplastics may be able to cross the blood-brain 
barrier (Kozlovskaya et al., 2014).

62

5.2.1 Research challenges

Factors that affect toxicity in humans may be indicated 
by  animal  studies  and  human  cell  studies,  but 
determining the relevance of lab findings to real life 
environmental exposure is now a research challenge. 
For  example,  surface  properties  of  nanoplastic 
particles  affect  their  absorption  in  biological  cells; 
eco-coronas  may  reduce  their  bio-reactivity  and 
potential  impacts  (Pulido-Reyes  et  al.,  2017;  see 
section 4.4); and negatively charged particles appear 
to  have  fewer  impacts  on  cell  functioning  (Xia  et 
al.,  2008;  Natarajan  et  al.,  2020;  Saavedra,  Stoll 
and  Slaveykova,  2019).  Aggregation  observed  with 
negatively charged nano-PS may also reduce uptake 
through membranes (Shen et al., 2019). Meanwhile, 
within organisms, protein coronas appear to promote 
translocation of nanoplastics (Shen et al., 2019).  

Biological  systems  do  not  interact  with  ‘bare’ 
nanoplastics,  but  a  complex  of  nanoparticles 
and  corona.  Addressing  the  knowledge  gap  on 
nanoplastic corona formation and cellular interaction 
is important to corroborate effects observed in vitro 
(Kihara et al. 2021). Researchers using nanoparticles 
designed  for  drug  delivery  often  use  nanoparticles 
in  the  bloodstream  ‘tagged’  with  opsonin  proteins, 
allowing  macrophages  (immune  cells)  to  recognise 
the nanoparticles and engulf them. This phagocytosis 
process is critical for eliminating pathogens and dead 
cells, and leads them to organs such as the liver (see 
section  5.1),  where  the  human  body  degrades  and 
clears  most  toxins.  It  has  been  estimated  that  30-
99%  of  administered  micro  and  nanoparticles  may 
accumulate in the liver (Zhang et al., 2016). 

Another  factor  affecting  movement  of  nanoplastics 
within  human  tissues  is  size.  In  one  study  cited  by 
Shen  et  al.  (2019),  for  example,  44  nm  PS  entered 
gastric  cancer  cells  at  a  higher  rate  than  larger  PS 
particles of 100 nm size. Smaller size may not always 
lead to higher uptake, though – PS nanoparticles of 
100 nm were taken up by human colon cancer cells 
more easily than 50 nm particles, in one study (Win 
and Feng, 2005). As well as size, the type of plastic 
material influences uptake of nanoplasutic particles 
by intestinal cells, for example, most likely influenced 
by differing lipophilicity (fat attracting) characteristics 
of  the  surface  as  it  interfaces  with  lipid  rich  cell 
mebranes (Paul et al., 2022).   

Fig. 22. Human immune system macrophages absorbing 
nanoplastic.  Macrophages  are  a  type  of  immune  cell 
which  find  and  clear  body-foreign  particles.  In  the 
human body, these "garbage collectors" are absorbing 
the  plastic  nanoparticles  and  recruit  further  immune 
cells  triggering  an  immune  response.  Bone  marrow 
derived  macrophages  were  placed  with  fluorescent 
nanoplastic particles (green) in the lab. After 24 hours 
the  cells  were  fixed  and  stained  –  blue  =  nucleus  of 
macrophage,  red  =  cytoskeleton  of  macrophage  and 
green  =  the  nanoplastic  particles.  The  NP  distorted 
the shape of the cell. Source: Marlene Schwarzfischer, 
University Hospital of Zurich. CC -BY-NC-ND

in  cell 

The  shape  of  micro  and  nanoparticles  is  also 
internalisation;  for  example, 
important 
filamentous  shapes  evade  phagocytes  better 
than  spheres,  and  rigid  particles  are  more  easily 
internalised than soft particles (Anselmo et al., 2015). 
Weber et al. (2022) found that irregular polystyrene 
triggered a higher inflammatory response compared 
to spherical nanoplastics. Characterising the shapes 
of  nanoplastics  to  which  humans  are  commonly 
exposed  is  probably  important  for  understanding 
health risk. 

63

Some  studies  suggest  plastic  composition,  and  the 
ability to carry contaminants is the most important 
factor  in  nanoplastic  toxicity.  The  inflammation 
themselves 
potential  of  nanoplastic  particles 
compared  with  nano  forms  of  inorganic  metal 
oxide  materials  is  negligible.  The  hazard  from 
nanoplastics  leaching  toxic  additives  has  caused 
more concern, as as well their ability to act as vectors 
for  attached  pollutants  (Mitrano  and  Wohlleben, 
2020). All these areas of research require attention 
to  inform  risk  assessment.  With  recent  research 
indicating that nanoplastics (Jeong et al., 2022), and 
plastic  monomers  may  accumulate  in  human  cells 
(Schwarzfischer  and  Rogler,  2022),  and  knowledge 
that other types of nanoparticles lead to pathological 
conditions (ibid.), the need is pressing.

Although polystyrene nanoparticles have been much 
studied,  relatively  little  attention  has  been  paid  to 
how they behave within cells, why cytotoxicity differs 
between  types  of  cell  (Jung  et  al.,  2020),  and  how 
cells may expel them on (Han et al., 2021). Studies 
investigating  the  behaviour  of  PS  nanoplastics  in 
mouse cells show that they accumulate in cytoplasm 
(as  they  cannot  be  broken  down  like  biodegradable 
pathogens),  causing  cellular  stress.  Indeed,  this 
presents  a  problem  for  nanoparticles  designed  to 
deliver drugs – they often meet a ‘dead end’ in the 
lysosome,  the  cell’s  ‘waste  compartment’  (Reinholz 
et  al,  2018).  Nanoparticles  can  also  be  exported 
from cells through exocytosis, but there is very little 
literature on the rate at which this occurs in human 
cells (Han et al., 2021; Han & Ryu, 2022). One study 
suggests  that  smaller  polystyrene  nanoparticles 
(50  nm)  exit  cells  more  easily  than  larger  particles 
(500  nm)  (Liu  et  al.,  2021).  This  study  also  noted 
that interactions with cells were dominated by forces 
other  than  electrostatic  charge  (the  hydrophobic 
effect and Van der Waals’ force).

5.2.1 Nanoplastics and the human body: state of science

While the discovery of microplastics in human lungs, 
blood  and  placentas  has  made  news  headlines 
(e.g.  Carrington,  2022),  evidence  of  nanoplastics 
in  sampled  human  tissue  is  yet  to  be  published. 
It  seems  inevitable  they  will  also  be  detected  in 

the near future. Meanwhile, understanding of their 
potential effects on human systems, and ability to 
translocate  in  the  body,  draw  on  in  vitro  studies 
with  human  cells  and  in  vivo  studies,  chiefly  on 
rodents.  

… information can be extracted from the fields of nanomedicine and 
nanotoxicology, where a wide variety of engineered nanoparticles were 
found to cross the epithelial barrier in both in vivo and in vitro models 

Rubio, Marcos and Hernández, 2019

 
64

Fig. 23. Studies on nanoplastic interactions with human tissues use human cells and animal models.  
Source: González-Acedo et al. (2021)

Since  ingestion  is  a  likely  possible  route  of 
human  exposure,  much  research  has  focused  on 
nanoplastic behaviour in the intestines. Ingestion 
can  lead  to  systemic  uptake  if  nanoplastics 
are  able  to  pass  through  the  intestinal  barrier 
(epithelium),  and  be  distributed  to  tissues  and 
organs  via  the  bloodstream  or  lymphatic  system. 
In a human intestine model, nano-PET was indeed 
able to cross the gut barrier, reported Magrì et al. 
(2018). Cortés et al. (2020) found PS nanoparticles 
within human intestinal cells exposed at a range of 
concentrations for 24 and 48 hours, but observed 
no significant toxic effects – similarly to findings of 
a different study exposing brain and epithelial cells 
to PS nanoparticles (Schirinzi et al., 2017). 

Some  studies  have 
investigated  effects  of 
microplastics on the microbiome (Lu et al., 2019), 
but few have used nanoplastics. Xiao et al. (2022) 
noted  changes 
intestinal  microbiota  after 
exposing mice to 50 nm PS for 30 days, as well as 
significantly  disrupted  intestinal  mucus  secretion, 
but no adverse effects on the liver, lungs or brain. 
It  is  worth  exploring  a  potential  link  between 

in 

nanoplastics  and  bowel  problems  –  one  study 
found higher levels of micro- and nanoplastics in 
faeces  of  participants  with  Inflammatory  Bowel 
Disease (Yan et al., 2022). 

Stem  cell-based  models  –  in  particular  organoids 
including  various  cells  of  the  epithelium  –  are 
emerging  that  may  offer  better  insights  than 
currently  used  models  based  on  cancer-derived 
cell lines (Bredeck et al., 2022). In a recent piece of 
research using human intestinal organoids, Hou et 
al. (2022) found distinct accumulation of 50 nm PS 
in  cells,  causing  inflammatory  responses  and  cell 
death. The degree to which ingested nanoplastics 
may  accumulate  in  human  tissue,  over  long 
timeframes, is a key area to be investigated through 
these new methods. Work by Mohamed Nor et al. 
(2021), for example, suggested microplastics may 
make up 0.004% of the mass of inorganic particles 
humans ingest each day, but how this translates to 
health risk is unknown. Interestingly, Tamargo et al. 
(2022)  provided  the  first  evidence  that  polymers 
may undergo biotransformation in the human gut, 
using simulated digestion of PET microplastics.

 
 
 
65

Fig. 24. Mohamed Nor et al. (2021) predicted lifetime exposure of children and adults to microplastics.  
Source: Mohamed Nor et al. (2021)

The  liver  is  one  of  the  first  organs  that  ingested 
nanoplastics  will  encounter.  Banerjee  et  al. 
(2022)  studied  the  interactions  of  50-500  nm 
PS  nanoparticles  and  human  liver  cells.  Smaller 
particles  were  more  readily  internalised  and  had 
greater toxic effects, however larger particles also 
induced  cell  death  and  inflammatory  response. 

The  researchers  emphasise  that  it  is  difficult  to 
determine  environmentally  relevant  doses,  and 
clearance  mechanisms  are  not  well  understood, 
so the length of time particles remain in systemic 
circulation  –  and  the  length  of  organ  exposure  –  
is not known. 

A comprehensive MP/NP dose response study using more exposure 
time points, particle sizes, various polymeric composition, and in cells 
from different organs is necessary to understand the impact of MP/NP 
exposure more clearly on human health  

Banerjee et al., 2022

Inhalation  of  aiborne  nanoplastics  must  also  occur. 
In  an  experiment  by  Xu  et  al.  (2019),  human  lung 
epithelial  cells  exposed  to  25  nm  and  70  nm  PS 
nanoplastics suffered negative effects; however, when 
Meindl et al. (2021) exposed human lung cells to 20 
and 200 nm PS nanoparticles for 28 days, they found 
that  healthy  respiratory  cells  adapted  to  low  levels 
of  repeated  exposure.  Rubio,  Marcos  and  Hernández 
(2019) assert that nanoplastics should be captured by 
mucus  in  the  respiratory  system;  whether  scientists 
will  find  nanoplastics  in  human  lung  samples,  as 
microplastics have been, remains to be seen, however 
Fournier et al. (2020) demonstrated in an animal study 
that maternal pulmonary exposure to nanoplastics led 
to nanoplastics in foetal tissues.  

Given the expected chronic exposure to nanoplastics in 
the environment, it is important to investigate whether 
they may accumulate in human cells. After exposing 

stem cells to PS nanoplastics for 48 hours, Jeong et al. 
(2022)  found  “excessive  amounts”  were  internalised, 
leading  to  growth  inhibition  and  reduced  renewal  of 
cells. They explained that exocytosis may have been 
hindered by the tightness of cell colonies. In terms of 
methodology,  they  emphasise  that  human  induced 
pluripotent  stem  cells  are  an  excellent  material  for 
investigating  long-term  toxicity  –  due  to  the  fact 
that  these  cells  self-renew  –  and  effects  on  foetal 
development.  In  contrast  to  many  similar  studies, 
they also tested control materials (silica nanoparticles 
and  PE  microplastics),  concluding  that  the  effect  of 
nanoplastics on human embryonic cells merits further 
investigation.  A  recent  review  of  in  vitro  intestinal 
models of nanoplastic health effects, suggested future 
hazard assessments could benefit from co cultures of 
more than one human cell line – as well as organ on 
a chip models –  as a more reliable insight into doses 
that trigger biological effects (Bredeck et al., 2022).

 
66

the large-scale intracellular accumulation … suggests the possibility of 
causing unexpected alterations in human embryogenesis ... 

Jeong et al., 2022

Other  areas  that  currently  being  investigated 
include  nanoplastic  interaction  with  elements  of 
the immune system – which act as first-line defence 
against foreign substances in the intestinal tract, 
lungs and liver, for example – and effects on cancer 
cells. Nano-PVC and PS have been found to induce 
inflammatory  cytokine  release 
in  monocytes 
(immune  cells  in  blood;  Weber  et  al.,  2022); 

another experiment showed nano-PS impaired lipid 
metabolism in macrophages (Florance et al., 2022) 
– but the big picture with regards to nanoplastics 
and immune function has yet to be seen. Likewise, 
some studies show that nano-PS may exacerbate 
cell  proliferation  in  cancers,  but  evidence  is  still 
being gathered (Barguilla et al., 2022).

Box 14: Does research reliably indicate human health risks? 

While  many  studies  are  now  being  published 
that  find  negative  impacts  on  human  cells  and 
animal  models  (Xu  et  al.  2022),  it  must  also  be 
noted that there is still much work to be done on 
reproducible findings and robust methodology. For 
example, concentrations and exposure times used 
in experiments vary greatly.

Leakage of fluorescent stains can distort results, as 
can  solvents  in  which  nanoparticles  are  dissolved 

(Stock et al., 2022). Bulky, hydrophobic flurophores 
themselves may influence in vivo fate of particles. 
Keinänen et al. (2022) have proposed radiolabelling 
of nanoplastics and positron emission tomography 
as  an  alternative  way  of  tracking  particles  in 
mammals. Their initial work found that the majority 
of  PS  nanoparticles  ingested  by  mice  were  not 
absorbed,  but  were  eliminated  within  48  hours. 
However, they only administered a single dose, which 
does not reflect real world, continuous exposure.

[N]anoplastics cannot be considered one homogenous entity when 
assessing their health implications and the use of spherical polystyrene 
nanoplastics may underestimate their inflammatory effects 

Weber et al., 2022

In  their  review  of  studies  on  the  human  health 
impacts  of  micro-  and  nanoplastics,  Xu  et  al. 
(2022)  found  that  only  16  of  133  studies  stated 
the  health  impacts  were  not  of  concern.  The 
consensus  is  that  nanoplastics  must  be  treated 
with  caution,  even  if  we  are  still  unsure  of  their 
effects on human health. 

Yan  et  al  (2022)  found  15  types  of  plastic  in 
faeces,  dominated  by  PET  and  polyacrylamide, 
yet  the  majority  of  toxicological  studies  still  use 
polystyrene  –  which  accounts  for  only  about  7% 
of  plastic  production.  Busch  et  al.  (2021)  found 
that  PVC  nanoplastic  was  more  toxic  to  inflamed 
liver  cells  than  PS  and  other  work  indicates  that 
PET,  nylon  and  polyacrylonitrile  had  different 
inflammatory  effects  compared  to  PS  (Busch  et 
al., 2022). More relevant sample material must be 
made available; for example, a new way to obtain 
PET nanoplastics from plastic bottles has recently 
been  presented  (Villacorta  et  al.,  2022).  Non-
spherical nanoplastics should also be studied.

 
Box 15: Nanoplastics and the brain

An  ongoing  EU  funded  study  called  ‘Nanoglia’  is 
examining  the  impact  that  nanoplastic  particles 
in  the  human  brain  may  have  on  neurological 
functioning. The researchers note that nanoplastics 
can  translocate  from  the  gut  to  the  lymph  and 
circulatory systems in mammals, also crossing the 
blood-brain barrier (EC, 2021). The Nanoglia project 
seeks to understand the impact of the nanoplastic 
particles  on  the  brain’s  immune  cells  –  microglia 
– which engulf these plastic particles. These brain 
cells are important for neuronal homeostasis, and 
may  be  activated  by  nanoplastic  reaching  the 
brain.  The  team  will  use  an  animal  brain  model, 
to  understand  how  this  activation  of  microglial 
cells  by  nanoplastics,  could  affect  foetal  brain 
development  –with  respect  to  potential  cellular, 
molecular and behavioural changes. 

It  has  been  shown  that  nanoplastics  cause 
behavioural disorders in fish (Sökmen et al., 2020), 
and  thus  may  also  represent  a  risk  for  human 
health,  in  particular  for  brain  function.  However, 
long-term  bioavailability  and  toxicity  of 
the 

67

nanoplastics  in  the  brain  are  unknown.  Microglia, 
as the main neuroimmune cells of the brain, have 
a defence function – needed during inflammatory 
conditions  –  but  also  constantly  sense,  and 
respond  to  environmental  changes,  to  maintain 
neuronal  homeostasis.  This  places  microglia  at 
the interface between normal and abnormal brain 
development,  and  function  –  shown  by  a  recent 
discovery that chronic microglial activation causes 
neurodegeneration  (EC,  2021).  When  microglia 
internalise  nanoplastics  in  the  brain,  this  might 
lead to their acute or chronic activation, triggering 
neurological disorders (ibid.).

The Nanoglia project team hope to understand how 
nanoplastics  trigger  microglial  activation  during 
embryogenesis  and  postnatal  stages  –  alongside 
establishing  whether 
immune  activation 
this 
leads to permanent changes in brain development 
and  function.  These  mechanistic  insights  into 
microglia  and  nanoplastics  in  rats  may  elucidate 
environmentally 
of 
neurological disorders in humans. 

pathogenesis 

triggered 

"Human brain digital illustration (©Shutterstock, photo by Yurchanka Siarhei)"

68

5.3 Risk assessment

As outlined above, research into potential human 
health  impacts  of  nanoplastics  is  underway. 
Despite inconclusive evidence of impacts in relation 
to  real  world  exposure,  and  in  light  of  concerns 
regarding toxic chemical additives leaching from 
ingested  nanoplastics,  many  believe  immediate 
action should be taken to ensure humans are not 
exposed  to  levels  of  nanoplastics  that  may  be 
hazardous to health. The extent to which micro- 
and nanoplastics are present in the environment, 
drinking  water  and  seafood,  however,  poses 
challenges  that  are  not  easily  addressed  by 
standard risk analysis (EASAC, 2020).

In nanoplastic risk assessment, it is key to consider:

 y Should we apply the Precautionary Principle?

An important issue is the ‘precautionary principle’, 
in relation to both negative effects of nanoplastics 
on  the  environment  and  potential  human  health 
the  precautionary 
impacts 
approach, 
is  based  on 
fragmentary  evidence  but  strong  indicators  of 

risk  management 

(Table  6).  Under 

negative  impacts.  Many  highlight  the  co-benefits 
of taking this approach, as summarised by Martin 
Wagner in his debate over the risks of microplastics 
with  Thomas  Backhaus:  “Based  on  my  values,  I 
favor  a  precautionary  approach  to  microplastics, 
not because I consider them doomsday devices but 
because  I  believe  in  positive  change”  (Wagner  & 
Backhaus, 2019).

 y Early action can be most cost effective

ECHA  (2020)  notes  that  as  microplastics  (and 
nanoplastics)  are  extremely  persistent,  based  on 
the  ‘option  value  theory  of  resource  economics’, 
it is appropriate to take cost effective action now 
–  despite  uncertainties.  The  costs  of  inaction  on 
nanoplastics require further exploration.

 y Heed early warning signs

The European Environment Agency proposed early 
warning signs, to be taken into account by regulators 
when considering materials and substances; these 
are highly relevant for the design and regulation of 
polymeric materials (Paulsen et al., 2021). 

69

Table 6. Comparing an evidence based and precautionary principle approach to microplastics, also applicable 
to nanoplastics. (Source: Backhaus and Wagner, 2018).

Arguments in 
favour

Strictly evidence-based approach

Precautionary approach

Insufficient knowledge 

Sufficient knowledge

 y Low exposure (on current 

estimates)

 y Low toxicity (on current 

knowledge)

 y Presence of natural particles at 

higher levels

 y Likelihood of negative impacts low

 y Ubiquity

 y Persistence

 y Mobility in food web

 y Increasing emissions

 y Part of microplastics problem where 

sufficient knowledge on impacts exists

 y Existence of unknown, negative 

impacts

Actions in 
favour

 y Identify knowledge gaps

 y Take risk decision 

 y Perform more research filling 

 y Develop and implement risk 

these gaps

 y Take risk decision

 y Depending on outcome: develop 
and implement risk management 
measures

management measures based on 
fragmentary knowledge 

 y Perform research into the 

effectiveness and efficiency of these 
measures

 y Refine measures

Advantages

 y Avoids inefficient risk 

 y Early action avoids negative 

management measures

impacts later

 y Avoids unnecessary opportunity 
and unintended externality costs

 y Motivates positive and economic 
change (vision of better society)

 y Avoids regrettable substitutions

 y Fosters technological and societal 

 y Reduces cost of action 

innovation

 y Reduces cost of inaction, 

induces change

Risks from plastic, including nanoplastic, are not only 
related to health. Marine litter also impacts negatively 
on  human  welfare,  affecting  fisheries,  tourism  and 
aquaculture, and often brings losses to communities. 
Smaller-scale  marine  plastic  litter  causes  chemical 
and  microbial  transfer  when  ingested,  acts  as 
a  transporter  of  biota,  and  can  change  species 
assemblages (JRC, 2016; EASAC, 2020). 

Nanoplastics  could  carry  chemicals 
into  the 
marine  environment,  as  well  as  adsorbing  trace 
contaminants  present  in  the  environment  onto 
their  surface  –  concentrating  them  –  so  they 

may  have  a  toxicological  impact  when  ingested.  
In  a  lab-based  study,  unidentified  components 
of  common  consumer  plastics  (PP,  LPDE,  PS  and 
PLA) had a number of toxic effects (EASAC, 2020). 
The  EASAC  report  (2020)  lists  906  chemicals 
‘probably’  associated  with  plastics,  including  63 
associated  with  human  health  hazards,  68  with 
environmental hazards, and seven classified in the 
EU as persistent, bio-accumulative and toxic. Frond 
et al. (2018) calculated that 20 of these chemicals 
would amount to 190 tonnes in the environment, 
and  that  microplastics  in  coastal  areas  (which 
degrade to nanoplastics) were associated with high 

70

level of PCBs in the marine environment – showing 
how small plastic particles can be a route of toxic 
chemicals  into  the  environment.  A  more  recent 
review  reportedmore  than  6  000  chemicals  in 
plastics, including 1 518 plastic-related chemicals 
of  concern,  which  the  authors  suggest  should  be 
substituted for safer alternatives (Aurisano, Weber 
and Fantke, 2021). The authors highlight the need 
for a global and overarching regulatory framework 
for plastics and related chemicals, in support of a 
circular economy for plastics and of target 12.4 of 
the UN Sustainable Development Goals.

The  research 
literature  has  established  the 
presence of microplastic particles in food destined 
for  human consumption, however, a similar level 
of  studies  investigating  nanoplastics  in  food 
items  is  not  yet  available.  Nanoplastic  particles 
have  been  found  in  some  crop  plants  grown  in 
polytunnels, as well as in fish and shellfish such as 
scallops  (Larue  et  al.,  2021;  Al-Sid-Cheikh  et  al., 
2018).  Additionally,  as  microplastic  particles  are 

known to degrade to nanoplastics when ingested 
by  earthworms,  krill  and  other  organisms  (see 
section  3.2),  it  is  not  unreasonable  to  conclude 
this might also occur in humans. Indeed, potential 
biotransformation of plastics in human digestion 
has recently been demonstrated (Tamargo et al., 
2022). Evidence on which to base a risk assessment 
of  human  health  impacts  of  nanoplastics  per  se 
is currently insufficient, but the potential impacts, 
combined  with  problems  related  to  macro-  and 
microplastic  pollution,  strongly  suggest  formal 
risk  assessment  for  nanoplastic  exposure  needs 
to  be  developed.  As  the  body  of  literature  on 
nanoplastic  toxicity  grows,  such  assessment 
will  in  future  be  able  to  draw  on  frameworks 
such  as  the  ‘adverse  outcome  pathways’  (AOP) 
approach to classifying evidence on the hazards 
of  substances  including  nanoplastics  (Jeong  & 
Choi, 2019; Hu & Palić, 2020).

To date, only four papers have been published about the estimation of 
human intake of microplastics, consistently concluding that the data 
available in the literature is hard to compare, incomplete and insufficient 
for a reliable assessment of MP ingestion. The assessment of the food 
safety risk related to MP/NP contamination is, consequently, impeded 

Vitali et al., 2022

5.4 Biomonitoring of plastic-associated chemicals and metabolites

Although  there  is  a  lack  of  evidence  upon  which 
to  formulate  a  risk  assessment  of  nanoplastics 
for  human  health,  research  has  found  chemicals 
associated  with  plastics  (including  nanoplastics) 
in the human body. Human biomonitoring involves 
measuring 
concentrations  of  environmental 
contaminants, and/or their metabolites, in human 
tissues  or  body  fluids,  such  as  blood,  amniotic 
fluid,  breast  milk,  saliva,  hair  or  urine  sources. 
Biomonitoring  is  the  gold  standard  in  assessing 
the  health  risks  of  environmental  exposures, 
because  it  provides  an  integrated  measure  of  an 
individual’s exposure to contaminants from many 
sources (Royal Society of Chemistry, 2019). 

For some chemicals, their widespread presence in 
the  general  population  at  concentrations  capable 
of causing harm – in animal models – has raised 
public health concerns. However, it should be noted 
that although biomonitoring is useful in identifying 
plastic-related  chemicals  and  their  metabolites 
in the human body, it cannot be inferred that the 
source  of  these  is  nanoplastics.  Additionally,  any 
data  for  regulatory  risk  assessment  must  meet 
quality standards.

Biomonitoring  has  proved  some  chemicals  used 
in  the  manufacture  of  plastics  are  present  in  the 
human  population.  For  example,  the  American, 
‘National Health and Nutrition Examination Survey’ 

 
(NHANES)23  is  one  of  the  most  comprehensive 
human  biomonitoring  programs  yet  undertaken. 
NHANES  reports  on  several  chemicals  associated 
with  the  use  or  production  of  plastics,  including 
bisphenol A (BPA), phthalates, styrene, acrylamide, 
triclosan  and  brominated  flame  retardants,  and 
their  concentrations  in  the  general  population.  In 
the EU, the H2020 European Human Biomonitoring 
Initiative (HBM4EU)24 coordinates work in this field 
to  provide  evidence  for  chemical  policy  making, 
with  bisphenols  and  phthalates  among  its  list  of 
priority substances. Studies under this initiative are 

5.5 Nanoplastics in the food chain

Microplastics  have  been  detected  in  beer,  honey, 
salt,  sugar,  fish,  shrimps  and  bivalves,  but  there 
is  currently  little  data  regarding  the  presence  of 
nanoplastics in samples of commercially available 
foods  on  the  market.  However,  nanoplastics  have 
been detected in lab-based studies of organisms, 
some of which are species which can be produced 
commercially  as  foods,  such  as  crop  plants, 
shellfish  and  fish  (Larue  et  al.,  2020;  Shen  et  al., 

71

ongoing, but results so far have reported average 
BPA  levels  of  1.78-1.97  µg/L  in  human  urine,  for 
example  (HBM4EU,  2020)  –  well  under  guidance 
values of 230 µg/L (adults) and 135 µg/L (children) 
urinary  total  BPA  derived  from  toxicological  data 
(Ougier  et  al.,  2021).  European  populations  are 
estimated to have average daily intakes of BPA of 
0.05 µg per kg of bodyweight – a value much lower 
that the tolerable daily intake set by EFSA, 4 µg per 
kg of bodyweight. The results do not indicate how 
much BPA exposure may derive from nanoplastics.

2019; Al-Sid-Cheikh et al., 2018). Meanwhile, data 
is  just  emerging  on  the  volume  of  nanoplastics 
that  may  be  released  from  products  that  come 
into contact with food and drink: Hernandez et al. 
(2019) found that plastic teabags release billions 
of nanoplastic particles into the drink; Zangmeister 
et al., (2022) report the release of trillions of sub-
100 nm particles from LDPE-lined cups.

A more exhaustive and trustworthy insight on MP and NP pollution along 
the food chain is fundamental to inform decision-makers towards the 
institution of strategies and infrastructures for the monitoring of those 
contaminants and for the setting of legal safety limits 

Vitali et al., 2022

In  the  study  looking  at  microplastics  in  beer, 
granular  material  aside  from  fibres  of  plastic 
and  microplastic  were  present,  but  couldn’t  be 
distinguished  as  either  sand  or  nanoplastics,  as 
spectroscopic analysis was not used in the study. It 
is likely – as microplastics degrade to nanoplastics 
–  that  both  types  of  particle  are  present  in  food 
that  are  consumed  by  humans  (Yee  et  al.,  2021; 
Liebezeit and Liebezeit, 2014).

In  a  study  using  Fourier  transform 
infrared 
spectroscopy on tap, bottled and spring water – all 
these sources were found to contain microplastic 
particles (Kosuth et al., 2018). Over 80% of samples 
of tap water collected from 159 worldwide sources 

contained  microplastics,  whilst  93%  of  samples, 
from 11 different brands of bottled water contained 
microplastics. From figures of average microplastic 
particles  in  different  food  and  drink  types,  it  has 
been estimated that humans consume around 39 
000  to  52  000  microplastic  particles  per  year  – 
more including inhalation of microplastic particles: 
74  000  to  121  000  particles  per  year  (Yee,  et 
al.,  2021).  Data  concerning  nanoplastics  are  not 
yet  available  mainly  due  to  a  lack  of  reliable 
quantification techniques (Yee, et al., 2021; EASAC, 
2020).  Yet,  it  seems  probable  that  they  will  be 
present at higher concentrations than microplastic 
particles in water and the air (Frehland et al., 2020).  

23  http://www.cdc.gov/nchs/nhanes.htm
24  https://www.hbm4eu.eu/

72

Vitali  et  al.  (2022)  examined  136  studies  on 
micro  and  nanoplastics  in  food,  including  marine 
organisms considered as food which were sampled 
in the field (Figure 25). Microplastic contamination 
of seafood was widely found, however they noted 
that about half of the studies examined mussels. 
Mussels  for  human  consumption  are  usually 
produced in dedicated farms and depurated before 
going  to  market,  therefore  wild  mussels  are  not 

relevant for food safety assessment. Although one 
study  identified  microplastic  in  chicken,  it  did  not 
prove presence in chicken muscles. The reviewers 
note that the assessment of micro and nanoplastic 
contamination 
in  animal  products  calls  for 
methods able to digest protein and fat rich animal 
tissues, and techniques for accurately quantifying 
particles small enough to pass the gut barrier into 
the animal’s edible tissues. 

Fig. 25.  Average chemical composition of microplastics isolated from (A) food, water and beverages; (B) 
seafood; (C) salt; and (D) water. CP, Cellophane; PA, polyamide; PEST, Polyester; PE, Polyethylene; PET, 
Polyethylene terephthalate; PP, Polypropylene; PS, Polystyrene; PU, Polyurethane; PVC, Polyvinylchloride; RY, 
Rayon. Source: Vitali et al. (2022)

Micro and nanoplastics in crops have received very 
little attention, though nanoplastics can penetrate 
roots, stems and leaves (see section 3.6). Of carrots, 
lettuce, broccoli, potato, apples and pears, sampled 
from  a  market  in  Catania,  Italy,  peeled  apples 
were  the  most  contaminated  with  microplastics, 
according  to  Conti  et  al  (2020)  –  however  there 
are  no  other  studies  with  comparable  data,  nor 
data  on  nanoplastics  in  crops.  Vitali  et  al.  (2022) 

therefore recommend a thorough investigation of 
the  nanoplastic  content  in  fruit  and  vegetables, 
to assess the food safety risk linked to crops and 
agricultural  soil  contamination.  However,  until 
reliable,  high-throughput  methods  for  analysing 
plastic  nanoparticles  are  available  –  along  with 
standardised,  comparable 
reporting  –  such 
assessment remains a challenge.

 
73

6. Summary

This  Future  Brief  presents  the  current  science  on 
nanoplastics:  their  detection,  assessment  and 
monitoring;  their  impacts  in  the  environment, 
ecotoxicity,  and  environmental  fate;  and  their 
potential  impacts  on  human  health.  In  2022, 
nanoplastics  research 
is  a  very  fast-moving 
area.  New  techniques  and  methods  are  currently 
in  development  to  detect,  identify  and  analyse 
nanoplastics and their impacts, in the environment 
and  in  organisms.  Methods  of  interest  currently 
include  various  advanced  techniques  in  light 
scattering and chemical identification. Harmonised 
protocols  and  reference  materials  are  needed  as 
well  as  advances  in  isolation,  identification  and 
field sampling. 

However,  there  is  much  we  do  know  already. 
Despite the challenges of measurement at a nano-
scale,  we  know  that  microplastics  degrade  into 
nanoplastics,  so  we  know  they  are  there,  even 
if  we  cannot  yet  detect  them.  There  is  abundant 
evidence  of  significant  detrimental  effects  of 
microplastics  –  and  of  plastics  at  any  scale,  with 
plastics and their additives being found frequently 
in the environment and in organisms. 

Plastic is now ubiquitous in the environment, and 
we,  and  Earth’s  other  lifeforms,  are  currently 
ingesting  and  inhaling  nanoplastics  at  unknown 
concentrations.  Evidence  has  emerged 
that 
nanoplastics are small enough to cross biological 
membranes, and to cross into the blood and brain, 
which  may  have  implications  for  human  health. 
The  effects  of  nanoplastics  once  they  enter 
organisms is less well studied – although there is 
some evidence of lethal and sub-lethal effects in 
various  species,  of  various  plastic  types,  at  both 
high concentrations and levels that may be found 
in  the  environment.  To  help  advance  research  in 
this area, projects under the European Cluster on 
Health Impacts of Micro- and Nanoplastics will be 
looking at nanoplastics’ impacts on health over the 
next few years.

It  is  also  understood  that  interaction  with  other 
organic  matter  in  the  environment  creates  ‘eco-
coronas’  –  organic  coatings,  which  can  modulate 
nanoplastics’  charge,  chemistry,  behaviour  and 
toxicity.  Since  there  are  still  challenges  when 
studying  nanoplastics  in  the  environment  (most 

research  has  been  conducted  in  the  laboratory), 
understanding  of  the  implications  of  eco-coronas 
for  nanoplastics  risk  assessment  in  naturalistic 
scenarios  is  as  yet  embryonic.  Science  is  only 
beginning  to  understand  where  nanoplastics 
are,  and  what  their  characteristics  are  –  and  is 
only  scratching  the  surface  of  understanding  of 
interactions at the nano-scale.

We also know that deliberate regulatory response 
is lagging behind the pace of nanoplastic release 
and  abundance  in  the  environment.  Recent  and 
ongoing  efforts  to  regulate  intentionally  added 
microplastics (i.e. ECHA opinion developed in 2019) 
have run into debate about the most appropriate 
lower  size  limit  for  classifying  microplastics. 
This  debate  has  been  mainly  inspired  by  the 
difficulty  of  enforcement  when  detection 
is 
so  challenging  at  smaller  sizes.  However, 
scientific  evidence  does  not  support  a  cut-
off  point  for  the  environmental  and  human 
effects of plastics at a particular size limit; 
indeed, while the relation between physico-
chemical properties and (adverse) biological 
effects appears complex and evades simple 
threshold  descriptions,  the  smaller  the 
plastic  particles,  the  more  likely  they  can 
cross  biological  membranes  and  the  more 
thoroughly they can permeate organisms. It 
follows, perhaps, that a combination of inclusive 
categorisation (where ‘microplastics’ also include 
‘nanoplastics’),  as  now  stands 
in  the  draft 
Commission proposal of 30 August 2022, placing 
the burden of responsibility on the producer, would 
avoid  leaving  a  gaping  hole  in  the  regulatory 
regime for intentionally added nanoplastics, and 
would reflect what we do know so far. 

However,  the  release  of  secondary  (‘unintentional’, 
degraded  plastics)  represents  by  far  the  greater 
volume  of  plastic  in  the  environment.  Most  of  the 
research  on  environmental  fate  of  micro-  and 
nanoplastics  has  focused  on  water  pollution. 
Pollution  of  soil  and  of  the  air  has  received  less 
attention,  and  so  these  are  areas  where  more 
research efforts could usefully be focused in coming 
years. The aggregation, concentrations, movement, 
bioavailability and bioaccumulation of nanoplastics, 
as  well  as  their  eventual  endpoints  and  lifecycle 
durations, most likely differ significantly according to 

74

the particular plastic material and its characteristics 
and additives. We also know that some plastics and 
plastic  additives  are  more  harmful  than  others. 
This gives one clear route for regulation: regulating 
on  the  basis  of  particular  plastics  and  materials, 
whatever their size. 

The  plastics  industry  is  a  valuable  part  of  the 
European  economy  with  a  complex,  international 
value  chain,  employing  1.5  million  people  and 
generating a turnover of €340 billion in 2015 (EC, 
2018).  However,  it  is  both  inevitable  and  urgent 
that  the  value  and  disposal  chains  of  plastic 
materials  are  significantly  rewritten  by  plastic 
producers  and  other  decision-makers  as  Europe 

transitions to a more circular and environmentally 
conscious  economy.  It  is  clear  from  the  research 
presented  here  that  the  full  lifecycle  of  plastics 
–  from  ‘cradle’  to  ‘grave’  –  is  not  complete  when 
we can no longer see the plastic: plastics continue 
to  have  environmental  effects  far  past  the  point 
they  become  invisible.  Failure  to  control  these 
invisible pollutants, which are already permeating 
the terrestrial, atmospheric, aquatic and biological 
environments, contributes to an escalating hazard, 
the  full  proportions  of  which  we  may  not  fully 
understand until it is too late. 

75

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Reviews, 203: 103118. 

 y Zhao, L., Qu, M., Wong, G. and Wang, D. (2017) 

Transgenerational toxicity of nanopolystyrene particles 
in the range of µg L− 1 in the nematode Caenorhabditis 
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 y Zhu, Z.L., Wang, S.C., Zhao, F.F., Wang, S.G., Liu, F.F. and 
Liu, G.Z. (2019) Joint toxicity of microplastics with 
triclosan to marine microalgae Skeletonema costatum. 
Environmental Pollution, 246: 509-517.

 y Zhu, H., Fu, S.F., Su, Y. and Zhang, Y. (2021) Effects of 
nanoplastics on microalgae and their trophic transfer 
along food chain: Recent advances and perspectives. 
Environmental Science: Processes & Impacts, 23: 1873-
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Wagner, M. (2019) Benchmarking the in vitro toxicity 
and chemical composition of plastic consumer products. 
Environmental Science & Technology, 53(19): 11467-
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 y Zimmermann, L., Bartosova, Z., Braun, K., Oehlmann, J., 
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GETTING IN TOUCH WITH THE EU

In person
All over the European Union there are hundreds of Europe Direct centres. You can find the address of the 
centre nearest you online (european-union.europa.eu/contact-eu/meet-us_en).

On the phone or in writing
Europe  Direct  is  a  service  that  answers  your  questions  about  the  European  Union.  You  can  contact  this 
service: 

—  by freephone: 00 800 6 7 8 9 10 11 (certain operators may charge for these calls), 
—  at the following standard number: +32 22999696, 
—  via the following form: european-union.europa.eu/contact-eu/write-us_en.

FINDING INFORMATION ABOUT THE EU

Online
Information  about  the  European  Union  in  all  the  official  languages  of  the  EU  is  available  on  the  Europa 
website (european-union.europa.eu).

EU publications
You can view or order EU publications at op.europa.eu/en/publications. Multiple copies of free publications 
can be obtained by contacting Europe Direct or your local documentation centre (european-union.europa.eu/
contact-eu/meet-us_en).

EU law and related documents
For access to legal information from the EU, including all EU law since 1951 in all the official language 
versions, go to EUR-Lex (eur-lex.europa.eu).

EU open data
The portal data.europa.eu provides access to open datasets from the EU institutions, bodies and agencies. 
These  can  be  downloaded  and  reused  for  free,  for  both  commercial  and  non-commercial  purposes.  
The portal also provides access to a wealth of datasets from European countries.

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