Analysis of currently available data for characterising the risk of engineered
nanomaterials to the environment and human health — Lessons learned from four
Karin Aschberger⁎, Christian Micheletti, Birgit Sokull-Klüttgen, Frans M. Christensen
European Commission, Joint Research Centre (JRC), Institute for Health and Consumer Protection (IHCP), Via E. Fermi, 2749, I-21027 Ispra, Italy
a b s t r a c t a r t i c l ei n f o
Available online 11 March 2011
Production volumes and the use of engineered nanomaterials in many innovative products are continuously
increasing, however little is known about their potential risk for the environment and human health.
We have reviewed publicly available hazard and exposure data for both, the environment and human health
and attempted to carry out a basic risk assessment appraisal for four types of nanomaterials: fullerenes,
carbon nanotubes, metals, and metal oxides (ENRHES project 20091). This paper presents a summary of the
results of the basic environmental and human health risk assessments of these case studies, highlighting the
cross cutting issues and conclusions about fate and behaviour, exposure, hazard and methodological
The risk assessment methodology being the basis for our case studies was that of a regulatory risk assessment
under REACH (ECHA, 20082), with modifications to adapt to the limited available data. If possible,
environmental no-effect concentrations and human no-effect levels were establishedfrom relevant studies by
applying assessment factors in line with the REACH guidance and compared to available exposure data to
discuss possible risks. When the data did not allow a quantitative assessment, the risk was assessed
qualitatively, e.g. for the environment by evaluating the information in the literature to describe the potential
to enter the environment and to reach the potential ecological targets.
Results indicate that the main risk for the environment is expected from metals and metal oxides, especially
for algae and Daphnia, due to exposure to both, particles and ions. The main risks for human health may arise
from chronic occupational inhalation exposure, especially during the activities of high particle release and
uncontrolled exposure. The information on consumer and environmental exposure of humans is too scarce to
attempt a quantitative risk characterisation.
It is recognised that the currently available database for both, hazard and exposure is limited and there are
high uncertainties in any conclusion on a possible risk. The results should therefore not be used for any
regulatory decision making. Likewise, it is recognised that the REACH guidance was developed without
considering the specific behaviour and the mode of action of nanomaterials and further work in the
generation of data but also in the development of methodologies is required.
© 2011 Elsevier Ltd. All rights reserved.
1.1. Regulatory aspects and risk assessment
Novel properties of materials that become evident within a size in
the nanorange (at least one dimension less than 100 nm) make them
attractive for exploitation in a wide spectrum of fields, including
information technology, energy production, environmental protec-
tion, biomedical applications, food, agriculture and many more. The
same distinctive chemical and physical properties of engineered
nanomaterials (ENM) that make them so attractive for new product
development have raised concern over their safety to health and
Environment International 37 (2011) 1143–1156
☆ The opinions expressed in this publication are those of the authors and not
necessarily those of the European Commission.
⁎ Corresponding author at: European Commission, DG Joint Research Centre (JRC),
Institute for Health and Consumer Protection (IHCP), Via E. Fermi, 2749, I-21027 Ispra
(VA), Italy. Tel.: +39 0332 789618; fax: +39 0332 785388.
E-mail address: firstname.lastname@example.org (K. Aschberger).
1Engineered Nanoparticles — Review of Health and Environmental Safety (ENRHES
project) 2009 EU 7th research framework programme http://ihcp.jrc.ec.europa.eu/
2ECHA (European Chemicals Agency), 2008. REACH Guidance on Information
Requirements and Chemicals Safety Assessment. http://guidance.echa.europa.eu/
0160-4120/$ – see front matter © 2011 Elsevier Ltd. All rights reserved.
Contents lists available at ScienceDirect
journal homepage: www.elsevier.com/locate/envint
The provisions of the current regulatory framework for chemical
risk assessment and management in the European Union, the REACH
regulation3(Registration, Evaluation, Authorisation and Restriction of
CHemicals) apply to engineered nanomaterials (EC, 2008a,b). How-
ever, the technical Guidance Document (ECHA, 2008) for preparing a
chemical safety assessment (CSA REACH terminology for a risk
assessment ultimately showinghow risks can be controlled) currently
include very little reference to substances in particulate and nano-
Within the EU funded project ENRHES (Engineered Nanoparticles:
Review of Health and Environmental Safety; Stone et al., 2009) a basic
risk assessment appraisal was carried out for four different classes of
nanomaterials, metals (with focus on nano-silver), metal oxides (with
focus on nano-titanium dioxide (TiO2) and nano-zinc oxide (ZnO;
environment only)),fullerenesandcarbonnanotubes (CNT),based on
available information in the literature. Each type of substance, e.g.
nano-silver, included different forms, e.g. differences in size and
shape, crystalline form, functionalisation and purity, and these
different forms can exhibit quite different properties.
The riskassessment appraisals wereintended to followthe general
approach specified in the REACH Guidance on Information Require-
ments and Chemicals Safety Assessment (ECHA, 2008) in a structure
similar to the format for preparing a Chemical Safety Report under
REACH, based on the four basic steps of a classical risk assessment:
hazard identification, hazard characterisation, exposure assessment
and risk characterisation. However, as the risk assessment for the four
case studies was built on publicly available data and only limited
information was available (specifically, a lack of knowledge on use,
exposure and risk management measures in place, as well as data on
inherent properties), the approach could not be followed in all details.
On basis of the identified information, the risk assessments for
human health and environment have been carried out following a
quantitative and/or a qualitative approach (see Fig. 1). The method-
ology for setting no-effect levels followed the relevant chapters of the
REACH guidance (Chapter R.8 and R.10 in ECHA, 2008). For the
environmental assessment, the quantitative approach requires the
determination of the Predicted Environmental Concentration (PEC)
and the Predicted No-Effect Concentration (PNEC) for each environ-
mental compartment (air, water, and soil). For human health, the
quantitative approach requires establishing exposure values for the
various routes of exposure (inhalation, dermal and oral) for
consumers and workers and the establishment of Derived No-Effect
Levels (DNELs), typically based on the extrapolation of animal data to
the human situation by using appropriate assessment factors.
To indicate the uncertainties associated with the derived values
and that they consequently cannot and should not be used for any
regulatory purposes we have decided to use the terms Indicative No-
Effect Concentrations (INECs) instead of PNEC and Indicative No-
Effect Levels (INELs) instead of DNELs. INECs have been compared to
PECs of different compartments and INELs to human exposure
estimates in order to identify potential risks. In case where no INEC
or INEL and/or no exposure values were available or could be
estimated, a qualitative risk assessment has been carried out. In
addition to the quantitative risk assessment, we have proposed a
qualitative ranking of the investigated ENM, based on exposure, use
and hazard information.
This current paper presents a summary of the results of the
environmental and human health risk assessment appraisals of the
four case studies in the ENRHES project, highlighting the cross cutting
issues and conclusions about fate and behaviour, exposure, hazard
and some methodological considerations. The risk assessment
appraisals were complemented with additional information, not yet
considered in the ENRHES project. It should be noted, that the rapid
increase in information of relevance for risk assessment of ENM may
3Regulation (EC) No 1907/2006 of the European Parliament and of the Council of 18
December 2006 concerning the Registration, Evaluation, Authorisation and Restriction
of Chemicals (REACH), establishing a European Chemicals Agency, amending Directive
1999/45/EC and repealing Council Regulation (EEC) No 793/93 and Commission
Regulation (EC) No 1488/94 as well as Council Directive 76/769/EEC and Commission
Directives 91/155/EEC, 93/67/EEC, 93/105/EC and 2000/21/EC.
PEC < INEC ?
Exposure < INEL?
Qualitative risk assessment
Indicative Human No-Effect
(decide on appropriate
Human exposure estimation
Decide on appropriate
Select dose descriptors from
(decide on threshold endpoints
Exposure assessmentHazard assessment
Risk adequately controlled?
Adequate data for risk
Fig. 1. Overall process flow of the risk characterisation procedure.
Adapted from technical guidance document; European Commission 2003 and REACH guidance; ECHA 2008.
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
quickly alter conclusions. Details of the individual case studies can be
found in the final report of the ENRHES project (Stone et al., 2009), in
Aschberger et al. (2010a,b) and in Christensen et al. (2010a,b).
2. Exposure assessment
2.1. Environmental exposure assessment
In light of the REACH requirements for exposure scenarios, the goal
of the environmental exposure assessment is to define production, use
and disposal conditionsguaranteeingthat risksfor theenvironmentare
controlled. This means that emissions to the environment should be
limited in each phase of the life cycle of ENM in such a way that the
resulting environmental concentrations are well below the PNEC.
To achieve this, it is essential to identify where ENM could end up
(e.g. water, sediments, and soil), in which form (e.g. aggregated and
degraded), at which concentrations, and which ecological compart-
ment and organisms could be exposed (e.g. benthic organisms, algae,
and soil organisms). Moreover, exposure concentrations in environ-
mental compartments have to be measured or estimated by models.
To the best of our knowledge, based in the literature identified,
there are no available data about actual environmental concentra-
tions, and the development of appropriate measurement techniques
is only at the beginning.
Some data are available for the release into the environment from
wastewater treatment plants. Farré et al. (2010) reported the results
of the determination of C60, C70, and functionalised C60fullerenes in
effluents of 22 wastewater treatment plants in Catalonia (Spain). In
half of the cases fullerenes were detected, thereof some in the μg/L
range (e.g. maximum C60concentration: 19 μg/L). Kiser et al. (2009)
studied the concentration of Ti in wastewater effluents from all
sources, reporting an average concentration of 16 μg/L of Ti. The study
results suggest that part of this Ti is in the form of nanosized oxides
from industrial/food origin.
These data can be used as input to estimate possible environmen-
tal concentrations. Some model estimates of ENM concentration in
environmental media were published in the literature (Boxall et al.,
2007; Mueller and Nowack, 2008; Gottschalk et al., 2009, 2010a,b;
Blaser et al., 2008). The environmental concentrations were derived
by using simple modelling approaches (EC, 2003; Boxall et al., 2007)
as well as a material flow analysis (Gottschalk et al., 2009, 2010a,b).
Concentrations were estimated for multi-walled carbon nanotubes
(MWCNT), nano-TiO2, nano-ZnO and nano-Ag in different regions of
the world (e.g. UK, Europe, and USA). However, both approaches were
largely based on assumptions, thus including a high uncertainty in the
modelled value. Moreover, the values estimated by different models
and in different studies for the same type of nanomaterial varied two
to four orders of magnitudes. Therefore, it was not possible to derive
PECs from the data available, and only the order of magnitude of the
estimated concentrations will be considered in this paper. Table 1
summarises the order of magnitudes of concentrations modelled and
measured in environmental compartments.
Fullerenes would be found mostly in soil and STP effluents, with
lower concentrations in surface waters. Carbon nanotubes are
expected to be found in sediments, with lower levels in soil and STP
effluents, and even lower in surface water. Nano-TiO2and nano-ZnO
are expected to be mostly found in soil and sediments, and less in
surface water. Finally, nano-Ag is expected to concentrate in
sediments and soils. From these data, it appears that apart from
surface water, also soils and sediments could be compartments of
concern due to the nanomaterials accumulation.
To complement quantitative estimations with their uncertainties, a
qualitative assessment of exposure evidences (i.e. amount produced,
type ofuses, emission data,and fate and transport)wascarried out.The
information was used in the risk assessment to justify an ENM ranking.
Very scarce and uncertain information is available for the production
volumes, generally referring to global scale and generally reported in
market study reports, or in companies and industry association's web
pages (Aitken et al., 2006; Gottschalk et al., 2010a). The highest
production was assumed for nano-TiO2(60,000 t/y), followed by nano-
ZnO (10,000 t/y), nano-silver and MWCNT (both around 500 t/y), and
C60(40 t/y). No data about the potential release into environmental
compartments during production are available to the authors.
The potential release into water, air or soil during consumer use
can be deducted from product category data. ENM contained in liquid
or powder products (e.g. nano-TiO2, nano-silver, and nano-ZnO) are
likely more easily released into the environment than ENM embedded
in solid matrices (e.g. CNT), as some studies showed (Brouwer, 2010).
In particular, the release of TiO2nanoparticles from house façades was
observed in a field study (Kaegi et al., 2008), and some lab
experiments suggested that a fraction (i.e. 4% to 10%) of silver in
textiles can be released during washing (Geranio et al., 2009; Benn
and Westerhoff, 2008) as nanoparticles. A study by Vorbau et al.
(2009) reported the release of particles from nano-ZnO coated
material, which was submitted to an abrasion process in a laboratory,
showing that the released ENM were embedded into the matrix
The environmental behaviour allows to estimate the fate of the
nanomaterials in the environment through properties such as
degradation, dispersion stability, solubility, and bioaccumulation. In
the following paragraphs, information and references relevant to
assess the environmental behaviour (e.g. degradation) of the case
studies are presented.
Although degradation and modification of the ENM coating and/or
functional groups is important because it may change the interaction
of the nanoparticle with the environment, only few studies were
published investigating this. Fortner et al. (2007) reported that C60is
oxidised in the presence of dissolved ozone in water, while Auffan
et al. (2010) reported the oxidation of polydimethylsiloxane used to
coat nano-TiO2as a dispersion adjuvant. In both cases, the increase of
dispersion stability in water was observed. Biodegradation of organic
coating of single-walled carbon nanotube by Daphnia magna was
observed by Roberts et al. (2007), decreasing the dispersion stability,
while complete degradation to CO2of fullerol (i.e. C60–OH) by white
rot fungi was observed by Schreiner et al. (2009).
2.1.2. Dispersion stability
Dispersion stability in solution is an important factor, since it
contributes to the transport potential and to the final fate of
nanomaterials. The main parameter is the agglomeration/aggregation
Orders of magnitude of modelled and estimated environmental concentrations of fullerenes, MWCNT, nano-TiO2, nano-ZnO, and nano-Ag.
Environmental compartment Fullerenes MWCNT Nano-TiO2
aSewage treatment plant.
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
state, which is influenced by a combination of many factors, i.e.
natural organic matter (NOM), colloidal clay, ionic strength (IS), pH,
and inherent ENM properties such as surface charge. In surface waters
at natural pH, low IS and high NOM lead to a stabilisation of the ENM
dispersion, and thus to long-range transport of the nanoparticles.
Relevant examples are the studies by Xie et al. (2008) on C60, Smith
et al. (2009) and Chappell et al. (2009) on carbon nanotubes,
Domingos et al. (2009) on nano-TiO2, Zhang et al. (2009) on nano-
TiO2and nano-ZnO, and Cumberland and Lead (2009) on nano-Ag.
These papers show that NOM stabilises the ENM in solution, and that
the presence of divalent cations increase aggregation. Also pH has a
role in the dispersion stability, influencing the zeta potential of
nanomaterials. It is relevant to highlight that according to Hyung et al.
(2007), MWCNT dispersed in natural waters are potentially stable for
long periods (at least 1 month). Domingos et al. (2009) reported that
nano-TiO2dispersion in natural surface water could be more stable
than expected from laboratory studies due to fulvic acids coating.
Battin et al. (2009) reported that nano-TiO2in natural waters can be
transported 5 to 10 km downstream before being completely
removed, with the bacterial community enhancing the nano-TiO2
sedimentation. Finally, Boncagni et al. (2009) reported that in
laboratory batch experiments, P25 nano-TiO2dispersion in deionised
water with NaOH was stable at pH 10 for at least 50 h.
The fate and transport in porous media (soil and sediments) are
much less studied, and only general considerations can be made.
Darlington et al. (2009) published a study about nanoparticle
characteristics affecting the transport through porous media, using
nano-Al as case study. They found that the agglomerate size, charge
and agglomeration rate in the transport medium are the most
important factors. The negative charge of the particle acted like in
water as a dispersion stabiliser. On the contrary to surface water,
according to Fang et al. (2009) NOM and clay in soil particles may
decrease the transport of nano-TiO2 nanoparticles in the porous
media. Finally, small pores increase the immobilisation of ENM (Jaisi
et al., 2008).
Solubility is relevant for metals and metal oxides, which could
leach soluble ions into water. Among the evaluated ENM, nano-ZnO
and nano-Ag are the most soluble. For nano-ZnO, the toxicity studies
highlighted that nanoparticles are almost completely soluble, and this
is important as their toxicity is related to ionic Zn (Zn2+) (e.g.
Heinlaan et al., 2008). According to the literature, the ion leaching of
nano-Ag is low (b4%), but still relevant to explain the potential
mechanisms of toxicity. However, the observed effect was suggested
to be nano-related (Fabrega et al., 2010). In particular, Navarro et al.
(2008) and Liu and Hurt (2010) suggested that the interaction of
nano-Ag with organisms could cause a localised Ag+ion release. In
addition, available studies indicate that nano-Ag solubilisation in
natural water may be lower than expected, due to the NOM coating of
particles (Gao et al., 2009).
To assess the uptake and bioaccumulation of nanoparticles by
organisms it is important to understand the mechanism of action of
nanomaterials and to forecast the transport along the food chain. C60
and CNT were observed to adhere to the external surfaces of
organisms such as fish embryos, algae, or plant roots (Blickley and
McClelland-Green, 2008; Wei et al., 2010; Cañas et al., 2008), or in the
external or internal (i.e. gut) epithelial cells of animals (Yang et al.,
2010). However, currently there are no evidences of CNT bioaccu-
mulating into tissues. Nano-Ag particles adhere to the external
surfaces of bacteria (Choi and Hu, 2008) or enter the fish embryos
(Laban et al., 2010). Kim et al. (2010) observed that nano-TiO2was
adsorbed to the body and to the antennae of D. magna. Pipan-Tkalec
et al. (2010) reported concerning terrestrial invertebrates that
uncoated ZnO nanoparticles lead to Zn bioaccumulation factors
similar to the macro-sized zinc. Limited uptake of nano-ZnO by
roots but no translocation in other tissues was observed in plants (Lin
and Xing, 2008).
2.1.5. Exposure profiles
Based on the available information only generic exposure profiles
for the investigated ENM can be drawn: Carbon-based nanomaterials
seemed to have a low exposure potential, since they are mostly used
as composite components. The release is less likely during use of
consumer products while during production and disposal, emission-
reduction measures are implementable. Carbon-based ENM are in
some cases degradable and not accumulated in biota. CNT dispersions
in surface freshwaters could be stable for long periods (up to
1 month) in the presence of NOM, allowing long-range transport
from the source. CNT transport in soil is depending, among other
factors, from the length of the tubes, and in general, they can travel for
short distances accumulating in the soil.
Nano-Ag is probably mainly released from consumer products
such as textiles, nano-Ag coated products and if used in liquids (e.g.
cleaning products). Nano-Ag is slightly soluble in natural water under
oxidising conditions, but according to the few data available, it may be
able to interact with organism as particle or small agglomerates
(especially if coated with NOM), and act as localised source of Ag+
ions. It is not possible to predict the dispersion stability in surface
water, since it depends on several environmental factors. However, in
freshwater transport for some distance is possible.
Nano-ZnO could be released by several consumer products such as
baby lotions and sunscreens, or from coated surfaces. Once in water,
uncoated nano-ZnO can be almost completely solubilised in water
releasing Zn ions (Zn2+). However, the effect of NOM stabilisation on
solubility and on particle transport range cannot be quantified. In
addition, the authors could not find anything about the behaviour of
coated nano-ZnO particles. Terrestrial organisms can ingest nano-
ZnO, which thus is acting as a source of zinc for the animals.
Nano-TiO2will be released by consumer products and by other
diffuse sources (e.g. facades). Nano-TiO2as such is not degradable, but
since it is normally coated, the degradation of the coating material
may modify properties such as stability in water. As for nano-ZnO,
NOM can enhance dispersion stability, increasing the transport
potential. Nano-TiO2uptake and bioaccumulation into organisms is
difficult to estimate due to lack of data.
2.2. Human exposure assessment
The greatest potential for human exposure is expected during
certain activities in occupational settings, where ‘raw’ nanomaterials
are handled in large quantities (Maynard and Aitken, 2007). Humans
can further be exposed during downstream occupational handling, as
consumers or via the environment. To our knowledge no information
has been published on the applicability to ENM of general exposure
estimation models, as e.g. those recommended in the REACH guidance
(ECHA, 2008) for estimating exposure to chemicals. Thus, the
2.2.1. Occupational exposure
Occupational exposure to ENM can occur via inhalation during
manufacturing, including bagging and handling and during formula-
tion into various preparations and during downstream activities.
Operators can be dermally exposed during the handling of powder or
liquid preparations of ENM. Ingestion can occur as a consequence of
swallowing of inhaledmaterialfollowingmucociliary clearanceorasa
result of hand-to-mouth contact.
For inhalation exposure estimation, it is important to capture all
relevant information on the amount (number, surface area or mass
concentration), size distribution, as well as shape, composition and
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
chemical reactivity of the ENM present in the air of the worker's
breathing zone. However, conventional methods are usually not well
suited to quantitatively measure and characterise ENM at the
workplace and usually rely on the mass of particles per unit volume
(Maynard and Aitken, 2007). Background nanoparticles can interfere
with quantitative exposure measurements and at present, there are
no rapid techniques available that would allow online distinction
between background nanoparticles and ENM (Ono-Ogasawara et al.,
2009). An accurate separation and distinction of ENM from the
background is extremely important for the evaluation of a possible
risk and for setting occupational exposure levels. Furthermore, the
ENM characteristics (physical changes and chemical composition)
may also change over time, due to ageing, agglomeration and
aggregation (Seipenbusch et al., 2008).
Discussions are ongoing regarding the best exposure metric(s)
(surface area, particle count/per particle size, particle mass, and particle
charge) for a correlation with observed toxic effects (see e.g. Tran et al.,
2000; Rushton et al., 2010). Surface area is recommended as a better
metric than mass concentration by some authors for rodent studies
(Oberdörster, 2001; Tran et al., 2000; Sager and Castranova, 2009),
in the workplace, and there is a need for appropriate practical
equipment, measurement protocols, and baseline measurements. In
general, there is a need to develop lower-cost, effective on-line
measurement devices for workplace measurement and the identifica-
tion of engineered airborne nanoparticles (Maynard and Aitken, 2007).
There have been recent efforts to improve the techniques to detect
the release of nanomaterials and characterise particles to render
routine measurement more specific and accurate, e.g. by portable,
direct reading (activity based) aerosol monitoring, combined with off-
line analysis of filter samples by microscopic and also chemical analysis
(Nanoparticle Emission Assessment Technique NEAT, Methner et al.,
2010a,b; Peters et al., 2009; Ono-Ogasawara et al., 2009).
Occupational inhalation exposure to specific nanomaterials has
been investigated and evaluated in a number of workplaces and pilot
plants/laboratories, in particular the latter (Maynard et al., 2004;
Demou et al., 2008; Yeganeh et al., 2008; Tsai et al., 2009). These
studies have reported a wide range of exposure values from different
activities and give mainly information on release sources and
potential exposure to ENM in the workplace and allow determining
whether engineering controls are effective in preventing exposure in
occupational settings. Considering that there are many limitations,
data from these different case studies were used as rough indications
for our risk assessment appraisal. It was the most practical way
forward to base our risk assessment on mass metrics, as this was the
metric generally available from exposure measurements and toxicity
Low exposure values ≤1 μg/m3were reported for CNT (e.g. 0.7 μg/
m3, from laser ablation facility; Maynard et al., 2004) and fullerenes
(Shinohara et al., 2009). Most reported exposure values were in the
range of ≈50 μg/m3, e.g. CNT (HiPCO process, Maynard et al., 2004),
non specific particle values for fullerenes (Yeganeh et al., 2008), or
estimated for 50 nm particles of nano-TiO2(althoughhighly uncertain
as based on ‘read-across’ from Aluminium oxide measurements in
Tsai et al., 2009). High exposure values ≥400 μg/m3were measured
for CNT in situations of high peak exposure (Han et al., 2008) or
without exposure control or estimated for 100 nm particles of nano-
TiO2(asabovebased on Tsai et al.,2009 using ‘read-across’). Fornano-
silver, one study reported peak concentrations of 7000 particles/cm3
(Tsai et al., 2009) and from another study worst case average
concentrations of about 60,000 particles/cm3(0.1888 mg/m3)
(Demou et al., 2008) were estimated. It should be noted that the
exposure levels ≤1 μg/m3are at the limit of detection and therefore
have to be taken with caution.
Dermal occupational exposure was only reported from one study
with SWCNT (Maynard et al., 2004). A dermal exposure value of
12 mg/person (both hands) not considering the use of gloves, which
are usually assumed to reduce the exposure by 90% (i.e. 1.2 mg/
person) was suggested for the risk assessment.
2.2.2. Consumer exposure
ENM can be found in a variety of consumer products and some
nanomaterials are already used for decades such as nano-silver as anti-
microbial, nano-TiO2in sunscreens or Carbon Black in ink and tyres.
to contain nanomaterials (see e.g. http://www.nanotechproject.org/,
Dekkers et al., 2007). Silver seems to be by far the mostly used
nanomaterial in consumer products.
Within the different inventories, the product categories ‘Home
furnishing and household products’ and ‘Health and Fitness’ contain
the majority of products (Wijnhoven et al., 2009). These categories
include the subcategories cleaning products, coatings, cosmetics,
clothing, personal care, sporting goods, sunscreen and water- and air-
filtration. The product group of highest production volume include
coatings and adhesives but also food packaging materials, catalytic
converters, automotive components and sun cosmetics. Based on an
expert panel (Wijnhoven et al., 2009) the highest possible exposure is
expected from personal care products and cosmetics (including sun
cosmetics, oral hygiene products, supplements and health products),
but also other products like fuel for motor vehicles (after combustion)
and do it yourself coatings, adhesives and cleaning products are
expected to lead to high potential exposure of consumers.
For many of the applications ENM are embedded into a matrix, e.g.
for improving surfaces, strength-weight properties or electrical
properties. The very limited information available on whether such
nanomaterials will stay fixed or could become available for exposure
to humans due to migration, evaporation, wash out or abrasion due to
wear and tear suggests that exposure is very low (Hsu and Chein,
2007). ENM used as surface coatings or enhancing materials in
consumer products (e.g. nano-silver in food containers or CNT in
textiles) may lead to oral or dermal exposure. ENM dissolved in liquid
suspensions, suchas paints (e.g.nano-TiO2), andcosmetics (e.g.nano-
TiO2, nano-ZnO and fullerenes) enable a direct (sometimes intended)
dermal contact or inhalation exposure if these products are used in
sprays (e.g. paints and sunscreen). Boxall et al. (2007) have estimated
potentially very high short term exposure levels if sunscreen sprays
are used indoor. Similarly, other spray applications may cause high
(peak) exposures to workers and consumers. Dermal exposure to
nanoparticles in sunscreens has been estimated by e.g. Hansen et al.
(2009) and Nohynek et al. (2010). Nano-silver is used in wound
dressings, which may cause dermal exposure and lead to silver uptake
(Vlachouet al., 2007). Some consumerproducts may also changetheir
exposure potential during the product life-cycle, e.g. for paints where
nanoparticles will be in liquid form when the paint is applied but in
solid form once the paint has dried (weathering and physical
A fully quantitative exposure assessment of consumers to
nanomaterials was not possible as the content of nano-sized materials
is only publicly available for a very limited number of consumer
products (Hansen et al., 2009). In general, further research on
consumer exposure, in particular for products causing direct expo-
sure, is warranted.
2.2.3. Exposure via the environment
Humans can be exposed to ENM via the environment followingthe
release during manufacture, downstream uses, incineration or wear
and tear of products. Some of the investigated ENM can also occur
naturally (e.g. fullerenes). Gottschalk et al. (2009) carried out a
probabilistic material flow analysis from a life-cycle perspective of
products containing engineered nanoparticles (CNT, fullerenes, nano-
silver, nano ZnO and nano-TiO2) for the US, Europe and Switzerland.
Predicted Environmental Concentrations (PECs) in the air (Europe)
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
were estimated to be b0.005 μg/m3for nano-TiO2, ZnO and fullerenes.
The lower and upper quantiles (Q(0.15) and Q(0.85)) were 0.006
and 0.02 μg/m3for nano-Ag and 0.0025 and 0.007 μg/m3for CNT,
respectively. It seems likely that the production volumes of ENM
will increase significantly in the coming years, increasing also the
3. Hazard assessment
3.1. Environmental hazard assessment
Environmental hazard assessment of ENM aims to estimate
Predicted No-Effect Concentrations (= PNEC, however in our case
called INEC) from ecotoxicity data for different environmental
compartments. Below we will briefly discuss the type of data available
for the different representative taxa in water, sediment, and soil. In
this section we report only the references deemed relevant to
highlight specific hazardous properties of the investigated ENM. For
a more detailed review see the ENRHES final report (Stone et al.,
It is difficult to reach a general conclusion about the toxicity of the
ENM investigated for the tested taxa, since data were derived for
different forms of the ‘same’ ENM (e.g. different size, crystal structure,
and surface coating), and frequently the ENM characterisation was
not carried out or reported poorly.
3.1.1. Aquatic environment (including sediment)
22.214.171.124. Fish. Most of the ecotoxicological studies published are acute
studies for fish in different life stages (e.g. embryos, juveniles, and
adults), while fewer papers report chronic studies. Carbon-based
nanomaterials are the most investigated, followed by nano-TiO2,
nano-ZnO, and nano-Ag.
The endpoints examined include mortality, as well as sublethal
effects such as development, growth, respiration, malformation,
oxidative stress, and gene expression. Results were sometimes
expressed as conventional toxicity endpoints (e.g. Lethal Concentra-
tion 50, LC50; No Observed Effect Concentration, NOEC), but often a
more generic result such as ‘no significant mortality’ was reported by
Carbon-based ENM were generally less toxic to fish than metal and
stress in liver and gills, as well as pathological liver effects. Lethal
effects were observed in only one short term study (C60in Danio rerio
embryos) (Usenko et al., 2008) at 0.2 mg/L C60dispersed in water
with solvents (i.e. DMSO). However, in other papers no mortality and
low sublethal effects in short term were observed up to 10 mg/L of
water-stirred C60(Blickley and McClelland-Green, 2008). Long term
studies carried out on SWCNT dispersed using a solvent (i.e. SDS)
showed sub-lethal effects (e.g. oxidative stress, brain swelling, and
increase of ventilation rate) in juvenile trout at 0.1 to 0.25 mg/L
(Smith et al., 2007). In general, lethal and sublethal effect concentra-
tions of carbon-based ENM ranged between hundreds of μg/L to some
Among metal and metal oxide ENM, nano-silver and nano-ZnO
were more toxic to different fish life stages than nano-TiO2. Observed
effects were death, respiratory stress and morphological defects
measured at two orders of magnitude lower concentrations than
nano-TiO2. Asharani et al. (2008) observed mortality of D. rerio
embryos at 25–50 mg/L in a short term study, but other studies
reported LC50down to 34 μg/L (Cheng et al., 2009) for Oryzias latipes
adults. Zhu et al. (2008) reported a LC50for D. rerio embryos and
larvae of 1.8 mg/L. The same study reported no mortality concentra-
tion for nano-TiO2at 500 mg/L.
126.96.36.199. Aquatic invertebrates. Crustaceans such as Daphnia or Tham-
nocephalus platyurus are generally tested as representatives for
aquatic invertebrates. Lethal and sublethal effects were assessed for
different life stages in short and long-term studies. Both, CNT and
fullerenes showed mortality at the same concentration levels. Lovern
and Klaper (2006) reported a LC50for D. magna offspring of 7.9 mg/L
for sonicated water dispersion, while Spohn et al. (2009) showed no
mortality at 24 mg/L for water-stirred dispersion. Gut tissues damage
caused by C60on D. magna were observed at 1 mg/L (Yang et al.,
2010). Kennedy et al. (2008) reported a MWCNTLC50for Ceriodaphnia
dubia of 50 mg/L.
Among metal and metal oxide ENM, nano-Ag and nano-ZnO were
highly toxic to crustaceans in acute tests, causing mortality at low
concentrations. Griffitt et al. (2008) reported an LC50for nano-Ag for
Daphnia pulex of 40 μg/L, while Heinlaan et al. (2008) observed an
LC50of 3.2 mg/L for nano-ZnO for D. magna offspring, and of 0.18 mg/L
for T. platyurus larvae.
Nano-TiO2caused mortality after short-term exposure (i.e. LC50
48 h) at concentrations between 2 mg/L (Zhu et al., 2010) and 20 mg/
L (Heinlaan et al., 2008) for D. magna neonates. However, Zhu et al.
(2010)foundthatlongerexposure(up to21 days)caused mortalityat
2 mg/L, and sublethal effects on reproduction at 0.1 mg/L.
188.8.131.52. Algae and aquatic plants. For metal and metal oxide ENM, algae
were the most sensitive aquatic organisms. No studies were found
testing C60toxicity, while one study showed cytotoxicity and growth
reduction of algae caused by large aggregates of oxidised MWCNT in
sea water at concentrations around 1 mg/L (Wei et al., 2010). Among
metal and metal oxide ENM, nano-ZnO was the most toxic, with
substantial growth reduction (EC50) at 42 μg/L (Aruoja et al., 2009) in
freshwater, while Wong et al. (2010) observed an EC50of 4.6 mg/L for
Thalassiosa pseudonana in seawater. In all available studies, it
was shown that the toxicity was mediated and caused only by leached
Zn2+ions. Ata similar concentrationlevel wasthetoxicity ofnano-Ag,
as reported by Griffitt et al. (2008) and by Navarro et al. (2008),
showing EC50 values of 190 and 9 μg/L, respectively. However,
differently than for nano-ZnO, for nano-Ag toxicity a nano-effect was
observed, even if mediated by leached ions. It was suggested that
particles are absorbed to algae, thus acting as localised source of toxic
ions (Navarro et al., 2008).
Nano-TiO2was less toxic, with an EC50between 5.8 mg/L (Aruoja
et al., 2009) and 14 mg/L (Hund-Rinke and Simon, 2006), which could
be caused by light reduction due to the aggregation of the TiO2
nanoparticles (Aruoja et al., 2009).
184.108.40.206. Sediment organisms. The limited data available showed that
CNT and C60 are not toxic (no lethal or sublethal effects were
observed) for sediment organisms. Oberdöster et al. (2006) observed
no toxicity of C60at 22.5 mg/L for copepods. Kennedy et al. (2008)
reported a LC50for copepods between 68 and 268 g/Kg of MWCNT in
No data were found for nano-Ag and nano-ZnO, while one study
investigated the effects of nano-TiO2 on a marine polychaete
(Galloway et al., 2010), showing sublethal effects caused by oxidative
stress at 1 g/kg.
3.1.2. Terrestrial environment
Information concerning the toxicity to organisms of the terrestrial
(e.g. nitrifying community), and crop species (e.g. spinach). A summary
of relevant information is reported in the following paragraphs.
220.127.116.11. Invertebrates and microorganisms. C60 was not toxic to
earthworms up to a concentration in food of 1 g/kg (Scott Fordsmand
et al., 2008), while CNT showed sublethal effects on reproduction
(long-term exposure) at 176 mg/kg in food (Scott-Fordsmand et al.,
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
2008). Studies carried out on bacteria and protozoan showed
ambiguous results, but no effects after long-term exposure were
observed at concentrations up to 50 mg/kg C60(Johansen et al., 2008).
Some studies were carried out on Porcellio scaber (terrestrial
isopod) and Eisena fetidafor nano-TiO2andnano-ZnO.No effectsup to
1 mg nano-TiO2/g animal wet weight were observed in isopods
(Drobne et al., 2009), while nano-TiO2and nano-ZnO were not toxic
to earthworms up to 1 g/kg (Hu et al., 2010). According to the
available data, effects on nitrifying bacteria were found at concentra-
tions below1 mg/L of nano-Ag. Nano-TiO2was nottoxic to bacteriaup
to a concentration of 100 mg/L (Velzeboer et al., 2008).
18.104.22.168. Plants. No data about effects of C60on plants were identified.
MWCNT were found not toxic or of low toxicity at concentrations
between 56 mg/L (Cañas et al., 2008) and 2 g/L (Lin and Xing, 2007).
For the other ENM, few studies were found, but the results concerning
germination, root growth and plant growth were ambiguous. Nano-
ZnO resulted in toxicity only for some species, with concentrations
between 20 mg/L (Lin and Xing, 2007) to 1 g/L (Stampoulis et al.,
2009) depending on the species tested. Nano-TiO2showed positive
effects on root growth and photosynthetic activity of Spinacia oleracea
up to 4 g/L (Zheng et al., 2005). Finally, nano-Ag was found to be toxic
to plants at concentrations between 50 mg/L (Kumari et al., 2009),
and 500 mg/L (Stampoulis et al., 2009).
3.2. Human health hazard assessment
Particle deposition in the respiratory tract is affected by particle
characteristics, including size, shape, and breathing rate (ICRP, 1994;
Borm et al., 2006). Nanoparticles readily reach the deeper lung, and the
alveolar regionis theprimarysiteof deposition for nanoparticle sizes in
the 10–100 nm range, whereas head and thoratic deposition predom-
inate for nanoparticles less than 10 nm. Poorly soluble particles (like
cleared due to macrophages actions (see e.g. Elgrabli et al., 2008). The
migration of MWCNT into the pleura has been demonstrated following
pulmonal exposure to high concentrations (Ryman-Rasmussen et al.,
pleura exists, through stomata and parietal pleura (Donaldson et al.,
2010), which may fail for long fibres, such as long MWCNT. More
knowledgeonthetransportto andresidence oflongfibres inthepleura
is of great importance, as the pleural cavity is the location in which
pathogenic fibres (such as asbestos) are known to elicit disease, such as
In general, limited information is available investigating the
systemic uptake following inhalation of ENM. Conflicting results
have been reported regarding translocation of particles from the lung
into the blood for several nanoparticles (reviewed by Borm and
Kreyling, 2004). It has been shown that silver is absorbed following
pulmonary exposure (see e.g. Takenaka et al., 2001; Sung et al., 2009)
however, it is still not clear whether the uptake occurs as nano-silver
particles and/or as silver ions. Following nasal exposure, some
ENM, like for example TiO2, have been shown to be absorbed and
accumulated in the brain, probably via neuronal transport (Wang
et al., 2008; Oberdörster et al., 2004).
Following oral exposure there was limited absorption of CNT
(Deng et al., 2007) and fullerenes (Yamago et al., 1995), whereas
some uptake has been demonstrated following nano-silver(Kim et al.,
2008) and nano-TiO2 (Wang et al., 2007) exposure. As for the
inhalation exposure, it is not clear whether the silver uptake occurs as
particle or silver ion.
The skin is usually considered a good barrier to prevent the uptake
of particles, even in nanoform, which can however be dependent on
different conditions, e.g. solvents (Xia et al., 2010) or flexing of skin
(Rouse et al., 2007). The available information suggest no or very
limited dermal penetration of fullerenes (Kato et al., 2009) and nano-
TiO2(see e.g. Johnston et al., 2009 for a review) through intact skin.
This may however need further investigations in particular for long-
term dermal exposure. Further investigations seems also relevant for
investigating the possible passage through damaged skin, especially
as some ENM are applied directly on sunburned or wounded skin,
such as may be the case for nano-TiO2in sunscreens and nano-ZnO in
ointments. Silver absorption has been shown following treatment
with nano-silver containing wound dressings (see e.g. Trop et al.,
2006; Vlachou et al., 2007). As for the other exposure routes it is not
clear whether silver is absorbed as particles or ions.
Once nanoparticles reach systemic circulation in the body, they
can be distributed to a number of different target organs including
liver, kidney, brain, lung, spleen and immunological system, where
they can accumulate and show signs of toxicity. However, most of the
available evidence on systemic distribution were obtained following
studies using internal (intraperitoneal and intravenous) exposure
(see e.g. Deng et al., 2007).
Based on the available evidence, it is not possible to make general
conclusions regarding the adsorption, distribution and longevity of ENM
in the body (Hagens et al., 2007). Even if data are available for one form
of a substance, it is difficult to generalise for that type of ENM as toxico-
kinetic behaviour is heavily influenced by physico-chemical properties
(e.g. size, surface charge, and functionalising groups). Distribution
studies with intravenous or intraperitoneal injection have indicated a
propensity for wide distribution of ENM to several organs, and the ENM's
toxicity has been investigated in several in vitro models. More knowledge
on the toxicokinetics is necessary to allow a better interpretation of
results gained with in vitro tests or following internal exposure to ENM.
22.214.171.124. Mechanisms of action. The main mechanism of ENM toxicity
seems to be oxidative stress, which triggers inflammation via the
activation of oxidative stress-responsive transcription factors. It has
been shown that chronic inflammation and oxidative stress caused by
ongoing exposure can lead to a number of particle-specific effects
such as fibrosis, genotoxicity and cancer caused by fibres or secondary
mutation. Physicochemical differences, such as the particle dimen-
sions and shape, purity, functionalisation etc. influence the type and
potency of the reaction; also dose, exposure time and the adminis-
tration type influence the results. More in-depth information on the
mode of action and the factors driving ENM toxicity of the four ENM
considered can be found in Johnston et al. (2009, 2010a,b,c).
126.96.36.199. Toxicity following oral exposure. The investigated ENM have
shown low acute and repeated dose oral toxicity. Signs of toxicity
were seen only with relatively high doses of nano-silver (Cha et al.,
2008) or nano-TiO2(Wang et al., 2007) applied.
188.8.131.52. Toxicity following dermal exposure. Dermal toxicity (argyria, a
skin discoloration, thus systemic toxicity (Trop et al., 2006)) was seen
following dermal application of wound dressings containing nano-
silver or unpurified SWCNT (inflammatory effects on skin, thus local/
topical toxicity (Murray et al., 2009)). The other tested CNTs,
fullerenes and nano-TiO2 did not show any signs of local dermal
effects, irritation or sensitisation on intact skin (see Aoshima et al.,
2009; Kishore et al., 2009 and Warheit et al., 2007). However, due to
the very limited information available to assess dermal toxicity, one
should be careful with drawing general conclusions. Further investi-
gations on the effects of ENM on damaged skin seem to be warranted.
184.108.40.206. Toxicity following inhalation/pulmonary exposure. The respira-
tory tract is considered a major portal of nanoparticle entry, because
of the likelihood that nanoparticles will become airborne during
handling. To investigate pulmonary effects, several studies have been
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
carried out, where ENM were administered via instillation, aspiration
and inhalation (e.g. Shvedova et al., 2005, 2008; Donaldson et al.,
2002, 2006; Lam et al., 2004; Ma-Hock et al., 2009; Pauluhn, 2010a).
The main effects typically seen are inflammation, granulomas and
fibrosis of the lung of different potency, depending on several
different factors such as dose, application, ENM type etc. Besides
local effects in the lung also systemic effects on the cardiovascular and
the immune system have been described (Mitchell et al., 2009; Lam
et al., 2006). In some cases, it was shown, that these systemic effects
were due to the release of pro-inflammatory signals and not
necessarily a consequence of systemic absorption (see e.g. Mitchell
et al., 2009). It is important to consider the possibility of such indirect
effects carefully whenevaluatingtoxicity tests for the risk assessment.
The mechanisms of ENM induced toxicity have been thoroughly
discussed by Johnston et al. (2009, 2010a,b,c).
severe effects than observed for inhalation (e.g. Lam et al., 2006), which
may be explained by the fact that a higher dose of particles reached the
lung via instillation. Results from bolus type dose delivery should
however not be used for purposes of risk assessment (Oberdörster,
2010) due to the uncertainty of the human relevance of the observed
effects and as it is difficult to ‘translate’ an instillation dose into an
Results from the inhalation studies are thus preferred/needed for
deriving human no-effect levels. However, as inhalation studies are
expensive and difficult to conduct, instillation studies are useful to
investigate or screen the potency and the relative toxicity of ENM.
For all four types of ENM considered in the case studies, dose
descriptors for risk assessment were identified for chronic inhalation
exposure. From the information identified, it appears that the effects
seen do have a threshold and therefore it seems appropriate to derive
INELs from the dose descriptors of the selected key studies (see
Table 2 for moredetails).Giventhe current knowledge, non-threshold
effects can however not be ruled out and the INELs derived are crude
estimates and should not be used for any regulatory decision making.
The selected key studies for the risk assessment appraisals are rat
inhalation studies of high reliability and sufficient nanomaterials
characterisation, where the animals were usually exposed for 6 h per
day and 5 days per week. For CNT we selected 2 recent 13-week
inhalation studies (OECD 413) with short MWCNT (b1 μm) in
compliance with good laboratory practice with Nanocyl 7000 (Ma-
Hock et al., 2009) and Baytubes® (Pauluhn, 2010a). The latter
included a post-exposure period for up to 6 months. The LO(A)EC
for Nanocyl was reported as 0.1 mg/m3(with minimal granulomatous
inflammation) while the same concentration was reported as a no-
observed adverse effect concentration (NOAEC) for Baytubes®. For
nano-TiO2(21 nm particles) we selected a 13-week inhalation study
(including assessment of pulmonary responses up to 52 weeks post-
exposure) (Bermudez et al., 2004) with a NOAEC of 0.5 mg/m3. For
silver a NOAEC and a lowest observed adverse effect concentration
(LOAEC) were determined from a 90 day inhalation study in rats
(Sung et al., 2008, 2009). At the lowest tested dose of 49 μg/m3(=
LOAEC) alterations in the lung function were observed, whereas other
effects, such as on the liver started at concentrations N133 mg/m3
which is considered a NOAEC for these effects.
Forfullerenes theNOAECwasdetermined at 2.22 mg/m3ina 10 day
study (3 h/day exposure) based on the absence of inflammatory effects
(Baker et al., 2008). A currently performed subchronic study with
fullerenes for which no results are available yet, suggests that the true
NOAEC even for subchronic exposure is much higher (Walker, 2009),
however this could not be considered for our assessment.
220.127.116.11. Genotoxicity and carcinogenicity. The genotoxic potential of
ENM is inconclusive to date as test results in vitro and in vivo seem to
depend not only on the material tested, but also on the experimental
set-up including the test system, exposure route, concentration
administered and the endpoint assessed (Johnston et al., 2009,
2010a,b,c). Genotoxic events can derive from the direct interactions
of particles with the DNA following cell internalisation, defined as
primary genotoxicity, or by the ability of particles to induce
inflammatory reaction and to generate an excess of ROS (secondary
genotoxicity). ROS can be generated by the ENM itself, by its metal
ions or impurities or by macrophages following a frustrated
phagocytosis of ENM such as aggregated, long CNT. Singh et al.
(2009) recently reviewed available evidence related to ENM geno-
toxicity and inter alia concluded that characterisation of the tested
form of ENM were too poor to draw any generalised conclusions.
Some ENM are suspected for being capable of inducing carcino-
genicity. For example, long straight MWCNT could be translocated
from the lung to the pleura where they could induce mesotheliomas,
similar to asbestos (Donaldson et al., 2010). A few studies using
intraperitoneal and intrascrotal exposure showed that long straight
MWCNT (N5 μm) could behave in a similar manner to asbestos and
have the potential to induce mesotheliomas in mice and rats (Poland
etal., 2008;Takagiet al.,2008;Sakamotoet al.,2009). Nocarcinogenic
effects were observed following intraperitoneal injection of short
MWCNT (b5 μm) (Muller et al., 2009), which could be explained by
the ability of phagocytes to ingest them and a mechanism of particle
clearance from the pleura, which may fail for long fibres (Donaldson
et al., 2010). It needs to be shown whether MWCNT have the ability to
induce mesotheliomas also in the pleural mesothelium through a
relevant exposure route, e.g. inhalation and this is a major issue for
the assessment of a carcinogenic risk of CNT.
Lung tumours have been reported following chronic inhalation of
very high doses (10 mg/m3) of nano-TiO2 (Heinrich et al., 1995).
chronic exposure of humans to low concentrations, as in the study very
high concentrations were used, which lead to lung overload in the
18.104.22.168. Reproductive toxicity. Toxic effects on reproduction and
development have been reported in vitro and in zebra fish for several
Overview of dose descriptors, overall assessment factors and estimated human indicative no-effect level (INELs) for workers of different ENM for chronic inhalation exposure.a
ENM type N(L)OAEC (μg/m3) Reference CommentModified N(L)OAEC (μg/m3) Overall assessment factorINELbworker(μg/m3)
Baker et al. (2008)
Bermudez et al. (2004)
NOAEC, 3 h/day, 10 days
NOAEC 90 days+52 w
NOAEC 90 days+90 days
LOAEC 90 days
LOAEC (reduced lung
function) 90 days
NOAEC (other effects) 90 days
MWCNTs100Pauluhn (2010a)50 252
Ma-Hock et al. (2009)
Sung et al. (2008)
133 Sung et al. (2009)671000.67
aThe numbers presented in this table are the results from the basic risk assessment exercise and should under no circumstances be used for any regulatory decision making.
bIndicative no-effect level.
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
ENM (e.g. Cheng et al., 2007, 2009; Komatsu et al., 2008). Nano-TiO2
(Shimizu et al., 2009) and fullerenes (Tsuchiya et al., 1996)
administered intravenously showed effects on the development of
the foetus, however it cannot be excluded that the effects are
secondary to accumulation of ENM in the placenta and insufficient
supply of the embryo/foetus. Also, given the current knowledge on
toxicokinetics, it can be speculated whether (and if so in which
quantities)ENM would reach this target following inhalation, oral and
dermal exposure. In conclusion, the relevance of these findings for
humans is questionable and further research needed.
3.3. Establishing indicative no-effect concentrations (INECs) and
indicative human no-effect levels (INELs)
3.3.1. Environment: INECs
For the environment hazard assessment, INECs were derived by
following the method to calculate the PNEC, thus using ecotoxicity
data and applying assessment factors (REACH guidance, ECHA, 2008).
Both acute and chronic toxicity data for representative taxa can be
used (e.g. fish, algae, sediment and terrestrial species, etc.). According
to the amount and type of information available, and depending on
the environmental compartment, the appropriate assessment factors
are chosen (10,000–1000 to 10-1). In general, the highest assessment
factor is used if only acute toxicity information is available, while
lower assessment factors can be used if studies provide chronic low or
no-effect concentrations on all required taxa or if a species sensitivity
distribution can be calculated.
Due to the uncertainty of the available ecotoxicological data, only the
For metal and metal oxide ENM INECs were derived only for the
freshwater compartment since insufficient data was available for the
other environmental compartments. The starting point was the lowest
assessment factor of 1000.
3.3.2. Human health: INELs
INELs were derivedonlyfor inhalationwhere the availabledata to
estimate no-effect levels and exposure levels was considered
sufficient. For dermal (except for one type of SWCNT) and oral risk
assessments, either exposure or toxicity data were lacking and thus
no INELs were derived and no risk characterisation could be
conducted. A major problem for deriving an INEL is the lack of
standardisation on how the nanomaterials in the study design are
Human no-effect levels, that are used in a risk assessment, are
generally based on dose descriptors (no observed adverse effect levels
or benchmark dose) for critical effects from animal toxicity studies
with modifications to the starting point and by applying assessment
factors (AF) (Chapter R.8 in ECHA, 2008). There have been no
epidemiological studies of workers involved with engineered nano-
particles (Schulte et al., 2009), which could allow or support a
quantitative risk assessment based on human data.
According to the guidance, the modifications to a human situation
take into account the differences of exposure conditions and include
the different exposure durations between laboratory animals (usually
6 h) and working hours (usually 8 h) and the different respiratory
volumes betweenrest and lightactivity (for workers). The assessment
factors for interspecies differences, intraspecies differences, extrapo-
lation for the duration, severity of the effect and confidence in the
database have to be applied, when considered appropriate.
In the conducted basic risk assessments, allometric scaling was not
applicable for interspecies variation, as the observed effects (local) do
not depend on the metabolic rate or systemic absorption and thus
only a factor (default 2.5) for other interspecies differences was
usually applied. For workers the default intraspecies factor of 5 was
applied and for the duration the default values for sub-acute to
chronic (6) and sub-chronic to chronic (2) extrapolation were
applied. Table 2 gives an overview on the selected dose descriptors,
the assessment factors and the estimated INELs. See Aschberger et al.
(2010a,b) and Christensen et al. (2010a,b) for more details.
4. Risk characterisation: results, interpretation and discussion
The goal of this environmental risk characterisation was to provide
a ranking of the investigated ENM on the basis of all available
information. The first evidence was the comparison between the
expected exposure and the order of magnitude of the INEC for metal
and metal oxides and hazard data for carbon-based ENM.
The INECs of nano-Ag and nano-ZnO were determined in the ng/L
range, while the INEC of nano-TiO2was in the μg/L range, thus three
orders of magnitude higher. For carbon-based ENM, not enough data
were available to estimate an INEC. Therefore, chronic low or no effect
levels in surface water were used as toxicity evidence in the
qualitative ranking by comparing their order of magnitude.
C60caused no or low effects to fish early life stages in concentra-
tions between μg/L and mg/L (Blickley and McClelland-Green, 2008),
while it showed no or low effects to invertebrates at concentrations in
the mg/L range (Yang et al., 2010) and no effects to soil organisms up
to g/kg concentrations (Scott Fordsmand et al., 2008). No effects
in freshwater organisms were seen with CNT in concentrations in the
soil organisms the order of magnitude was between mg/kg and g/kg
(e.g. Lin and Xing, 2007). The comparison to available exposure data
highlights that the expected environmental concentrations are three
to six orders of magnitude lower than the hazard data (i.e. no or low
chronic effect concentrations), both in water and in soil.
For nano-TiO2, both, the predicted exposure and effect concentra-
tions were only partially overlapping, with expected exposure
concentrations (i.e. ng/L) lower than the effect concentrations (i.e.
μg/L), while for nano-ZnO and nano-Ag, the predicted exposure and
the INEC were both in the ng/L range. However, due to the uncertainty
in the exposure estimation, and taking into account the worst case
approach, the PEC/INEC ratio resulted to be higher for nano-ZnO
(e.g.≈1000) than for nano-Ag and nano-TiO2 (i.e.≈1). Further
information on PEC/INEC ratios, which would allow to further
differentiate the potential adverse effects of the investigated ENM
can be found in the final ENRHES report (Stone et al., 2009).
A qualitative risk characterisation based on the available informa-
tion about exposure, fate and transport, and toxicity supported the
conclusionof a higherriskof metal andmetal oxide ENM.In general,it
was considered that high production volumes and the use in
consumer products as liquid or powder increased the exposure
potential. Severe effects (such as mortality) and in general adverse
effects starting already at low concentrations indicate a high (eco)
toxicity. A combination of these considerations was used to define a
ranking based on our judgment.
The considerations supporting the conclusion are briefly described
below and a more detailed discussion of the methodology and results
can be found in the final ENRHES report (Stone et al., 2009). The final
qualitative ranking can be summarised as follows.
Nano−ZnONNNano−AgNNano−TiO2NMWCNT ¼ C60
Of highest concern seems nano-ZnO, as it is used in several
consumer products in liquid form (e.g. body lotion, baby creams, and
sunscreens), leading to a higher exposure level. Moreover, it is highly
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
toxic (μg/Lrange) andinhibitinggrowthof primaryproducers(algae).
Several (but not all) laboratory studies indicate that the effects of
nano-ZnO are exclusively ion mediated (i.e. effect due to Zn2+
leaching). However, the uncertainty of the risk estimation for nano-
ZnO is high, and more studies are needed that investigate the
solubility of nano-ZnO in commercial form (e.g. coated ZnO) in the
Next we ranked nano-Ag, which is used in a wide range of products
with a high possibility of release into the environment (e.g. textiles and
coated surfaces). It has been demonstrated that nano-Ag can be released
from textiles in the form of particles, and that in the natural aquatic
environment not all nano-Ag is solubilised in oxidised form (Ag+). In
fact, several studies reported a distinct particle-related ecotoxic effect on
lower food chain levels (e.g. bacteria and algae). Even if for nano-Ag a
concern because of its increasing use and due to uncertainties in the
mechanism of action (i.e. ion or particle related) and environmental
behaviour (e.g. solubility and uptake of particles).
Nano-TiO2is considered of moderate concern compared to nano-
ZnO and nano-Ag, mostly because of its low ecotoxicity (LC50is three
orders of magnitude higher), even if a PEC/INEC ratio was comparable
to that of nano-Ag. As nano-TiO2 is used in several commercial
products, and it has the highest production volume among the ENM
investigated in this study, it can reach the environment through many
diffuse sources and the concern is mainly based on the potential high
Finally, solely based on qualitative data, the concern for C60and
MWCNT is expected to be low in comparison to the other investigated
EMN. The exposure potential is considered low, due to the currently
low production volume of fullerenes and due to the expected little
release of CNT from materials where it is embedded into solid
matrices (e.g. conductive plastics). The potential release from the
disposalsitesis in principlecontrolled. The ecotoxicity ofboth carbon-
based ENM was low compared to other ENM, showing only sub-lethal
effects at high concentrations in both freshwater and soil organisms.
The uncertainty was considered higher for C60than for MWCNT, due
to lack of information on the type of products and volumes produced.
The data available so far are not sufficient to perform an absolute
environmental risk assessment, especially due to the lack of both
exposure and chronic effect data. It can however be concluded that risk
is mainly expected from nano-Ag and nano-ZnO. In surface water risk is
mainly expected for algae and invertebrates with potential issues of long-
term effects on populations due to the impact on organisms in the
beginning of the food chain and predator's early life stages. Risk in soil is
not expected at the actual estimated environmental concentrations.
However, the exposure of soil organisms is expected to increase, due to
the low degradability and low mobility of ENM. In particular, the use of
sludgeto fertilise soils will representa diffuse source of ENM, with the risk
of accumulation over time. A similar issue may be raised for sediments,
where deposition of non-degradable ENM into sediments, especially at
river estuaries, will lead to potentially high exposure of filter-feeders to
ENM. Finally, there are not enough data to discuss the long-term toxicity
of realistic concentrations of ENM in the environment.
4.2. Human health
4.2.1. Risk characterisation
The INELs for CNT, fullerenes and nano-TiO2were calculated to be
between 1 and 17 μg/m3while those for silver were below 1 μg/m3
(Table 2). Although fullerenes are considered being low toxic and the
NOAEC was much higher than for CNT or nano-TiO2, the derived INEL
was lower than for nano-TiO2. This results from relatively high
assessment factors due to the short duration of the selected key study
and still high uncertainty due to only limited data availability. The
results may change, when more results from sub-chronic guideline
studies become available.
Human health risks are mainly expected from long term inhalation
of ENM at the workplace during activities of high release and when
exposure is not controlled. Our risk assessment appraisals showed that
INELs of nano-TiO2, CNT or fullerenes are higher or in the range of the
lower determined exposure values (≤1 μg/m3), while those for silver
were lower. All INELs were however lower than most of the reported
exposure values (medium≈50 or high N400 μg/m3) and therefore it
be excluded. However, no definitive conclusions can be drawn given
the scattered information available (see Aschberger et al., 2010a,b and
Christensen et al., 2010a,b for further details). It should also be noted
that the derived INELs are with in a range close to the limit of detection
for ENM in the air. This would make an exposure control at such low
exposure concentrations extremely difficult.
The modelled environmental concentrations for air (in the ng/m3
range) are several orders lower than the derived INELs for the general
public (usually by a factor of 2 lower than worker INELs, thus around
1 μg/m3, except silver, where it would be around 0.1 μg/m3) and
therefore based on these data no risk is currently expected. However,
the production level and consequently the environmental exposure
may increase in the future.
Semi-quantitative risk characterisations of other exposure routes
(dermal and oral) and populations (consumers) were hampered by a
severe lack of data, although risks were with some exception
generally considered low. Direct exposure to consumer products,
such as sprays or dermal application needs further attention.
In general, limited exposure information is available to date for
occupational exposure and almost nothing available for consumer
exposure and possible exposure via the environment. Finally, we
haven't identified any information on analysis or validation of the
applicability of existing human exposure models.
4.2.2. Relative toxicity of the investigated ENM
A comparisonof the toxicity of the differentENM is difficult as only
the data from one form of each of the different ENM substances were
used to derive the INEL, and this may not be representative for all
forms of an ENM substance. However, we have attempted to compare
the different toxicities and conclude that from the investigated ENM,
nano-silver is probably the most toxic with the lowest LOAEC, the
highest assessment factor proposed and consequently the lowest
INEL, below 1 μg/m3. The higher toxicity of silver probably results
fromthe release of silver ions from thesurface of nano-silver particles.
In contrast to that, nano-TiO2is considered to have a relatively inert
surface and to be less toxic. Comparable to that Rushton et al. (2010)
have assigned in a hazard ranking based on in vitro/in vivo activity for
inflammatory potential nano-TiO2to a hazard category ‘very low to
low’ — depending on crystalline state and nano-Ag to a hazard
category ‘high’. For the carbon based nanoparticles, CNT and full-
erenes, it is difficult to make any comparison to other ENM, as their
toxicity is heavily influenced by the shape (CNT), functionalisation or
the content of (metal) impurities. Also the size (and surface area)
needs to be taken into account when comparing toxicity between
different ENM. See e.g. Duffin et al. (2007) for further details on
4.2.3. Considerations of the data and the methodology used in the human
health risk assessment
The above-described methodology for establishing INELs followed
the REACH guidance on information requirements and chemical
safety assessment (Chapter R.8 in ECHA, 2008) and applied default
modification and assessment factors to the dose descriptors for
deriving the INELs.
The assessment factors used for deriving the INELs might not be
specific enough to consider the nanoparticle specific properties, as the
behaviour in the respiratory tract (e.g. deposition, persistence and
clearance) may be different as compared to those assumed for
K. Aschberger et al. / Environment International 37 (2011) 1143–1156
‘normal’ chemicals. The mode of action/toxic effects may also show
more differences between animals and humans as compared to what
is assumed for the ‘normal’ chemicals. Therefore, different assessment
factors when assessing ENM might be considered to account for the
differences between laboratory animals and humans, including
pulmonary deposition, retention half time, alveolar deposition,
alveolar macrophage volume or surface area of pulmonary alveoli
(see e.g. Oller and Oberdörster, 2010 or derivation of occupational
exposure limit by Pauluhn, 2010b, and NEDO risk assessment on CNT
by Kobayashi et al., 2009). A comparison with these other methods to
derive ‘safeexposurelevels’ showsthatthe approachusing thedefault
values from the REACH guidance is more conservative. The applica-
bility of these assessment factors for nanoparticle effects should be
reconsidered and it is up for discussion whether it is possible to
generate assessment factors for ENM based on the differences in the
parameters mentioned above. Further research and stakeholder
dialogue on the interspecies assessment factor are needed.
Further, it can be reconsidered whether the default assessment
factor for inter-individual differences of 5 for workers and 10 for the
general public could be changed, if it is proven, that differences
between individuals to show adverse effects following nanoparticle
exposures are different than those following exposure to ‘normal’
chemicals.Basedon ourevaluation of the literature,we however donot
see any clear evidence that this is the case. It should be noted, that the
REACH guidance default assessment factors should generally represent
a conservative approach but can be replaced by substance specific
values, if sufficient information and a scientific justification is provided.
In the EU, the European Commission are currently conducting two
REACH Implementation Projects onnanomaterials(so-calledRIP-oN 2
and 3), which aim at developing advice on how the REACH guidance
on information requirement and chemical safety assessment (ECHA,
2008) could be updated to address the specific properties of
nanomaterials. These projects include both human health and
environment. Results are expected in the first half of 2011.
Results indicate that the main risk for the environment is expected
from metals and metal oxide, especially for algae and Daphnia, due to
exposure to both, particles and ions. The main risk for human health
may arise from chronic occupational inhalation exposure, especially
during activities of high particles release and uncontrolled exposure.
Overall, due to the limited available data for performing a risk
assessment and due to the uncertainties in relation to the suitabilityof
the available test methods and risk assessment methodology it is
currently difficult to draw any firm conclusions for engineered
nanomaterials based on the open literature. Thus currently any
evaluation can only be made on a case-by-case basis (as is also
recommended by SCENIHR, 2009) and representative testing or
generalisations are not even possible within one ENM substance type.
consequent investigations of the case studies, that more data are
needed for a better hazard and exposure assessment. Toxic effects of
different ENM seem to depend on the size and shape (e.g. length of
ENM) and physico-chemical properties (metal content, aggregation/
agglomeration, surface chemistry, and functionalisation). It is pres-
ently unclear how well standard (eco)toxicity tests, designed for
soluble chemicals can be used to assess the (eco)toxicity of
nanomaterials (SCENIHR, 2007), and in any case harmonised disper-
sion and other dosimetry protocols have to be agreed. It is also unclear
whether each species of ENM or agglomerate of ENM may need
specific safety tests (and risk assessment), because of the (eco)
surface characteristics, amount and kind of impurities; most of these
characteristics being inherent to the respective specific production
As a conclusion from our study the following activities are
− Generation of reliable exposure scenarios, including data about
potential sources of ENM (production, ENM form, use category,
and emission factors) fate and transport in the environment (e.g.
degradation, bioaccumulation, solubility, and final environmental
compartment), as well as measurement of environmental
− Generation of reliable occupational and consumer exposure data
identification and quantification of ENM and the distinction from the
background. Beside point estimations also distributions and proba-
size distributions and possible agglomeration should be dealt with.
− Proper characterisation and agreement on proper metric(s) of
ENM to allow relevant comparison of toxicological dose–response
and exposure data.
− Further investigation of ecotoxicity, including long-term studies,
especially of algae and aquatic invertebrates, which seem to be the
most sensitive species. At the same time, the mechanism(s) of
action should be assessed both at the cellular and the genetic level.
More studies for the terrestrial compartment are required, due to
the use of STP sludge (relevant sink of ENM from treated waters)
for soil fertilisation.
− For human health further investigations into ENM toxicokinetics
are required, particularly the generation of data to provide clarity
on which ENM become systemically available followinginhalation,
dermal and oral exposure.
− Further investigations of systemic (cardiovascular and immuno-
logical) effects and sustainability and reversibility of effects
following repeated (inhalation) exposure, also considering,
whether systemic effects could be triggered via the release of
mediators; post-exposure observations periods (of several
months) should be considered.
− Further in vitro and in vivo studies to determine primary and/or
secondary genotoxic effects and investigations of the conditions
leading to positive and negative results.
− Depending on the above and possible further screening, consider
for some ENM the conduction of chronic/carcinogenicity studies
via inhalation to assess a possible risk (and the conditions) for
carcinogenic effects in exposed humans.
for the generation of information on exposure and toxicity, also the
risk assessment methodology can be adapted, including what
assessment factors would be appropriate for ENM.
Declaration of interest
The author's affiliation is as shown on the cover page. The authors
have sole responsibility for the writing and content of the paper.
We acknowledge the financial support of the EU 7th research
framework programme via the ‘ENRHES’ project.
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