L
`
Fate of Manufactured Nanomaterials in the
Australian Environment
G.E. Batley and M.J. McLaughlin
CSIRO Niche Manufacturing Flagship Report
March 2010
Prepared for the Department of the Environment, Water, Heritage and
the Arts
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necessarily reflect those of the Australian Government or the Minister for the Environment,
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While reasonable efforts have been made to ensure that the contents of this publication are
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Fate of Manufactured Nanomaterials in the Australian Environment
EXECUTIVE SUMMARY
With growing production and use of manufactured nanoparticles in a large range of consumer
products, regulatory agencies worldwide are addressing the risk that these substances may
pose to both the environment and human health. An assessment with respect to ecosystem
health requires an ecological risk assessment that must take into account current knowledge
about nanomaterial uses, environmental concentrations, fate, and effects, to determine both
predicted environmental concentrations (PECs) and predicted no-effect concentrations
(PNECs).
This report reviews the available literature on the fate of manufactured nanomaterials in the
aquatic and terrestrial environment. Seven classes of nanomaterials were considered: (i)
metal oxides; (ii) carbon products (n-C60 fullerenes, carbon nanotubes); (iii) metals; (iv)
quantum dots and semiconductors; (v) nanoclays, (vi) dendrimers, and (vii) nanoemulsions.
The key processes that govern nanoparticle behaviour in the aquatic environment are
aggregation and dissolution, driven by size and surface properties of the materials. These
processes can be mediated by interactions with dissolved organic matter and other natural
colloids. Biological degradation processes and abiotic degradation via hydrolysis and
photolysis do not appear to be significant in waters, although oxidation/reduction reactions
can be significant for some metals.
Similar processes are operative in terrestrial systems, but mobility is much reduced compared
to aquatic environments. Interactions of nanoparticles with soil minerals and organic matter
have not been evaluated, but are likely to be a function of particle size, shape and surface
properties (specific surface area and surface charge). Small hydrophilic nanoparticles (<20
nm) with net negative surface charges are likely to be mobile, while large hydrophilic
positively-charged particles will be sorbed by soil. Strongly hydrophobic nanoparticles are
likely to be strongly retained by soil organic matter.
Many parallels can be seen in the behaviour of natural colloids. In considering the behaviour
of manufactured nanomaterials, it is important that studies be carried out in natural waters as
the often orders of magnitude higher concentration of natural colloids can have a significant
impact. Aggregation results in growth of nanoparticles, often to sizes in excess of the
nanoparticle size definition of <100 nm, ultimately leading to sedimentation. This growth can
be prevented by the presence of surfactants and other surface coatings, or through the presence
of natural humic materials. Fibrillar colloids enhance precipitation. Any toxicity studies will
need to separately address particular nanomaterial formulations.
There is evidence to suggest that the impact of manufactured nanoparticles on aquatic
organisms differs compared to their macroparticle equivalents. In some instances such
assessments can be confounded due to nanoparticle solubility, with zinc oxide and cadmiumcontaining quantum dots being cases in point.
Mechanisms of nanomaterial toxicity include cellular damage due to oxidative stress, physical
damage to the cell surface, dissolution at the cell surface, and impacts via bioaccumulation.
The latter involves interaction with the cell surface for unicellular organisms and uptake
Fate of Manufactured Nanomaterials in the Australian Environment
iii
across the gill and other external surface epithelia for higher organisms. Bioaccumulation via
the food chain is also possible. The extensive literature on bacterial toxicity was considered
inappropriate for defining no effects concentrations of nanomaterials in waters, however, the
effects, both positive and negative, of some nanomaterials on bacteria present in sewage
treatment plants and on soil microorganisms have been discussed.
There are limited data on toxicity of nanoparticles to algae, invertebrates and fish. In the case
of n-C60 fullerenes, toxicity was highly dependent on the method of preparation, with the
particles dispersed by evaporation of tetrahydrofuran (THF) extracts being more toxic than
those dispersed by sonication, due supposedly to secondary effects of THF. Insufficient data
were available to derive high reliability environmental guidelines, but for freshwaters, low
reliability guidelines were derived for n-C60 and TiO2 nanoparticles. The calculated PNEC
values are only marginally above the concentrations estimated to be released to the
environment in calculations based on nanomaterial usage in the UK.
There is much less information on the behaviour and toxicity of nanoparticles in terrestrial
systems, due to difficulties in assessing dose against a background of natural nanoparticles in
the soil matrix. Heterogeneity and incorporation of nanoparticles into soil is also an issue for
ecotoxicological testing. There are a few reports of adverse effects of some nanoparticles to
terrestrial species cultured in vitro, but to date there is no strong evidence that nanomaterials
have significant adverse effects on terrestrial species in soil exposures. Further studies are
needed with a wide range of terrestrial species, and a wide range of nanoparticulate materials
in a range of soil environments, to determine if the preliminary data are sound.
There are numerous examples to demonstrate that nanomaterials can be bioaccumulated by
organisms. The extent to which this uptake exerts toxicity is less certain.
Current international activities in relation to nanoparticle risk assessment are discussed. In
summary, most countries see the need for more data gathering and research to improve the
risk assessment of these materials. This review indicates that the same is possibly true for
Australia, but the way ahead is reasonably clear. The basic recommendations for future
research are:
1. There is a need for measurements in natural water, sediment and soil samples of the
stability, and short- and long-term fate of the various likely formulations that might
reach these compartments of the environment. As well, techniques are needed to
distinguish natural from manufactured nanoparticles. These measurements should
focus on particle concentration, size and surface characteristics (area and charge).
2. Toxicity testing needs to be undertaken on nanoparticle formulations assessed in (1)
above. The tests should involve at least five species from four trophic levels as
required to derive PNECs using species sensitivity distributions. It is critical that
appropriate verification of particle and solute dose be undertaken in all ecotoxicity
testing, necessitating significant effort in (1) above.
3. As a precursor to toxicity testing, it will be necessary to develop standard (and valid)
methodologies for the hazard ranking of nanomaterial toxicity. These will need to
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Fate of Manufactured Nanomaterials in the Australian Environment
ensure the stability of the nanoparticle suspensions over the duration of the
standardised toxicity tests.
4. Comparisons of toxicity testing in natural vs. synthetic soil and water samples
demonstrating the effects of natural colloids.
Understanding the fate of nanoparticles in the Australian environment will assist risk assessments by
guiding the toxicity testing of nanomaterial formulations under real environmental conditions,
yielding realistic PNECs. This should be coupled with the development of appropriate
measurement techniques that can quantify both concentrations and particle sizes with
appropriate quality assurance and quality control. As well as size and composition, it is
evident that surface properties of nanoparticles will be fundamental in determining fate and
toxicity in the environment and these properties will need to be considered in any hazard
ranking.
A check list has been provided to incorporate fate considerations in assessing both
environmental exposure and effects of manufactured nanomaterials.
Fate of Manufactured Nanomaterials in the Australian Environment
v
Contents
EXECUTIVE SUMMARY ............................................................................................. iii
1.
INTRODUCTION ................................................................................................. 1
2.
CLASSES OF NANOMATERIALS ...................................................................... 2
3.
NANOPARTICLE USAGE IN AUSTRALIA ......................................................... 7
4.
CHARACTERISTICS OF ENVIRONMENTAL NANOPARTICLES ...................... 9
4.1
Manufactured Nanomaterials .................................................................................... 9
4.2
Natural Nanoparticles ................................................................................................ 9
5.
ENVIRONMENTAL SOURCES OF MANUFACTURED NANOPARTICLES ..... 12
6.
FATE OF NANOMATERIALS IN AQUATIC SYSTEMS .................................... 14
6.1
Key Pathways .......................................................................................................... 14
6.2
Behaviour of Manufactured Nanoparticles .............................................................. 15
6.3
7.
Aggregation ......................................................................................................... 15
6.2.2
Nanoparticle Solubility ........................................................................................ 17
6.2.3
Role of Nanomaterial Formulations and Impurities ............................................. 19
6.2.4
Fate in Natural Water Systems ........................................................................... 20
6.2.5
Nanoparticles as Vectors for Contaminant Transport ......................................... 21
Fate of Manufactured Nanomaterials in Terrestrial Systems .................................. 22
6.3.1
Key Pathways ..................................................................................................... 22
6.3.2
Behaviour of Natural Colloids in Soils ................................................................. 23
6.3.3
Behaviour of Manufactured Nanoparticles in Soils ............................................. 24
ECOLOGICAL RISK ASSESSMENT OF MANUFACTURED NANOPARTICLES
.......................................................................................................................... 25
7.1
8.
6.2.1
Polymeric Nanoparticles as a Separate Class ........................................................ 26
EXPOSURE ASSESSMENT ............................................................................. 26
8.1
8.2
What to Measure..................................................................................................... 26
Methods for Measurement of Nanoparticles ........................................................... 27
8.2.1
8.3
9.
vi
Relevance of OECD Test Guidelines .................................................................. 28
Modelling Exposure ................................................................................................. 29
ECOTOXICOLOGY OF NANOPARTICLES ..................................................... 33
9.1
Ecotoxicity and Nanoparticle Dose Metrics ............................................................. 33
9.2
Toxicity to Aquatic Biota .......................................................................................... 35
9.2.1
Mechanisms of Biological Uptake and Toxicity ................................................... 35
9.2.2
Ecotoxicity to Individual Species ......................................................................... 36
9.2.3
Developing Appropriate Guidelines for Nanomaterials in Waters ....................... 43
9.2.4
Bioaccumulation .................................................................................................. 44
9.2.5
Ecological Impacts .............................................................................................. 45
Fate of Manufactured Nanomaterials in the Australian Environment
10.
9.3
Sediment Toxicity.................................................................................................... 46
9.4
Toxicity to Terrestrial Biota ..................................................................................... 46
9.4.1
Ecotoxicity to Individual Species ......................................................................... 46
9.4.2
Development of Guidelines for Nanomaterials in Soils ....................................... 49
INTERNATIONAL PROGRESS ON NANOMATERIAL RISK ASSESSMENT .. 49
10.1 International Approaches ........................................................................................ 49
10.1.1
USA .................................................................................................................... 49
10.1.2
United Kingdom .................................................................................................. 51
10.1.3
Other International Activities ............................................................................... 52
10.2 Australian Activities ................................................................................................. 53
11.
DEVELOPMENT OF TECHNICAL GUIDELINES FOR NANOMATERIAL
ASSESSMENT.................................................................................................. 54
11.1 Exposure Assessment Incorporating Nanomaterial Fate ....................................... 55
11.2 Effects Assessment Incorporating Nanomaterial Fate ........................................... 56
11.3 Possible Approaches to Environmental Hazard Ranking of Nanomaterials ........... 58
12.
RESEARCH NEEDS ......................................................................................... 59
13.
ACKNOWLEDGEMENTS ................................................................................. 59
14.
REFERENCES.................................................................................................. 59
15.
GLOSSARY ...................................................................................................... 73
Fate of Manufactured Nanomaterials in the Environment
vii
List of Figures
Figure 1. Structures of (a) fullerene and (b) single-walled and (c) multi-walled carbon nanotubes
Figure 2. Schematic representation of mechanisms whereby surfactants help disperse SWCNTs. Top –
SWCNT encapsulated in a cylindrical surfactant micelle, middle – hemi-micellular adsorption of
surfactants on SWCNTs, and bottom – random adsorption of surfactants on SWCNT (from Ke and Qiao,
2007)
Figure 3. Typical structure of a dendrimer (first-generation polyphenylene dendrimer reported by Müllen
and coworkers in Chem.-Eur. J., 2002, 3858-3864).
Figure 4. Major types of aggregates formed in the three-colloidal component system: fulvic compounds
(or aggregated refractory organic material), small points; inorganic colloids, circles; rigid biopolymers,
lines. Both fulvics and polysaccharides can also form gels, which are represented here as gray areas into
which inorganic colloids can be embedded. (From Buffle et al., 1998)
Figure 5. Potential sources of manufactured nanoparticles to the environment
Figure 6. Pathways for manufactured metal oxide nanoparticles in natural waters
Figure 7. Electron micrographs illustrating aggregation of zinc oxide nanoparticles from dispersion of a
ZnO nanopowder (nominally 30 nm) in a freshwater algal medium, pH 7.5
Figure 8. Illustration of the solubility of amorphous silica as a function of radius of curvature (adapted from
Bjorn et al., 2006)
Figure 9. Key processes in soil relating to transformation and potential risk from manufactured
nanoparticulate particles
Figure 10. Framework for deriving mass flow data for silver flows from biocidal plastics and textiles (from
Blaser et al., 2008).
List of Tables
Table 1. Classes of manufactured nanomaterials
Table 2. Usage of nanomaterials in the commercial sector in Australia
Table 3. Aggregation data for manufactured nanomaterials in water (adapted from Boxall et al., 2007)
Table 4. Predicted environmental concentrations of manufactured nanoparticles in UK soil and waters
(from Boxall et al., 2007)
Table 5. Comparison of UK exposure data for manufactured nanoparticles with toxicity data (from Boxall
et al., 2007)
Table 6. Predicted environmental concentrations (PEC) of nano-Ag, nano-TiO2 and CNTs in air, water
and soil. (RE: realistic scenario; HE: high emission scenario) (from Mueller and Nowack, 2008)
Table 7. Hazard quotients (PEC/PNEC) for nano-Ag, nano-TiO2 and CNT in water (RE: realistic scenario;
HE: high emission scenario) (from Mueller and Nowack, 2008)
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Fate of Manufactured Nanomaterials in the Australian Environment
Table 8. Approach to toxicity testing of nanomaterials in waters
Table 9. Summary of toxicity testing results for manufactured nanomaterials (expanded from Apte et al.,
2008)
Table 10. Data for estimation of guideline concentrations for n-C60 in freshwater
Table 11. Published evidence of nanoparticle uptake by aquatic organisms (from Apte et al., 2008)
Table 12. Toxic effects of nanomaterials on soil organisms (from Klaine et al., 2008)
Fate of Manufactured Nanomaterials in the Australian Environment
ix
1.
INTRODUCTION
The last decade has seen an amazing growth in nanoscale science and technology, to the
extent that nanomaterials are now a component of a wide range of manufactured
products, from sunscreens to sensors. Given that the production volumes of some of
these materials is already exceeding thousands of tonnes, there is growing public and
regulatory concern for the potential adverse effects that release of these materials to the
environment may have on both human and ecosystem health. Nanoparticles are already
present in our environment in large quantities, derived both from natural sources
(volcanic dusts or natural bushfire products in air, colloids in aquatic systems and soils),
and as a consequence of anthropogenic activities (e.g. smoking, motor vehicle exhausts,
industrial stack emissions). Nanoparticles in the size range 3–7 nm have been shown to
account for more than 36–44% of the total number of particles in some urban air
samples, with the total size of particles ranging from <10 to 10,000 nm(Shi et al., 2001).
The adverse effects on human health of such nanoparticles in the atmospheric
environment (usually referred to as fine and ultrafine particles) have been well studied
and there are clear concerns for the finer particles that can reach the deeper recesses of
the lungs. For terrestrial and aquatic environments, there has been extensive research on
natural colloids (Buffle and Leppard, 1995a), however, there have been few studies of
anthropogenic particles.
Manufactured nanomaterials can be defined as those that are deliberately produced
rather than materials that are by-products of other activities not targeted at nanomaterial
production. Nanomaterials are commonly based on nanoparticles, for which the
accepted definition is particles that have at least one dimension less than 100 nm, but
the term is also used to refer to materials such as surfaces with nanometre-sized features
that are not particulate in nature, or substances with nanometre size voids. Small size
gives materials properties that differ from those of bulk or macroscopic materials. In
particular, optical, electrical and magnetic properties can differ in ways that are subject
to the laws of quantum rather than classical physics. Nanoparticles have a large surface
to volume (and mass) ratio, and potentially greater reactivity and mobility. Surface
areas can be as high as 1000 m2/g, far higher than conventional catalysts for example.
They have the tendency to agglomerate into larger microparticles, losing their distinctive
nano properties, although manufacturers are devising coatings that can stabilise
nanoparticles. Smaller size carries with it the potential to be more bioavailable, able to
penetrate biological membranes or to enter cells by endocytosis (engulfing by the cell
wall).
This report reviews the current state of knowledge with respect to nanoparticle fate and
effects in the environment, with a particular focus on aquatic and terrestrial systems, to
provide a foundation for the risk assessment of manufactured nanoparticles in Australia.
Fate of Manufactured Nanomaterials in the Australian Environment
1
2.
CLASSES OF NANOMATERIALS
Manufactured nanomaterials currently fall into one of at least seven different classes, as
shown in Table 1. The first class comprising metal oxides are common in their bulk,
non-nanoparticulate forms, and they are now being produced in nanosized forms that
capitalise on their enhanced properties. A case in point is zinc oxide that has been used
for many years as an opaque sunscreen because of its UV-absorbing properties,
scattering light in the range 200–700 nm.
Table 1. Classes of manufactured nanomaterials
Class
Metal oxides
Carbon
products
Component
Use
Zinc oxide
Cosmetic sunscreens and UV coatings; paints, plastics
and packaging
Titanium dioxide
Cosmetics
Cerium dioxide
Automobile catalyst
Mixed oxides
Cosmetics
Fullerenes
Plastics, catalysts, battery and fuel cell electrodes,
super-capacitors, water purification systems,
cosmetics, orthopedic implants, conductive coatings,
adhesives and composites, sensors, and components in
electronics, aircraft, aerospace and automotive
industries
Single-walled and
multi-walled
carbon nanotubes
Amorphous
carbon
Inks, photocopier toner, automobile tyres
Silver
Bactericide in wound dressings, socks and other
textiles, air filters, toothpaste, baby products, vacuum
cleaners, and washing machines
Iron
Remediation of groundwaters, sediments, soils
Gold
Electronics in flexible conducting inks or films, and as
catalysts
Bimetallic
nanoparticles FePd, Fe-Ni, Fe-Ag
Remediation of organics in waters; usually supported
nanoparticles
Quantum dots
and
semiconductors
CdTe, CdSe/ZnS,
CdSe, PbSe and
InP
Medical applications, photovoltaics, security inks, and
photonics and telecommunications
Nanoclays
Hectorites,
layered double
hydroxides
Cosmetics, toothpaste, antacids, paint additives,
catalyst supports, flame retardants, drug delivery
agents
Metals
Dendrimers
2
Coloured glasses, chemical sensors, and modified
Fate of Manufactured Nanomaterials in the Australian Environment
electrodes; in medicine as DNA transfecting agents,
therapeutic agents for prion diseases, formation of
hydrogels, drug delivery, DNA chips, and ex vivo
amplification of human blood cells
Emulsions
Acrylic latex and
other
formulations
Paints and surface coatings; sunscreens and similar
cream formulations; in medicine for drug delivery;
pesticide formulations
In a nanoparticulate form, ZnO is transparent to the eye, but retains much of its ability to
absorb UV radiation, albeit over a narrower spectrum. In Australia in 2005, of the 1200
sunscreens authorised by the Therapeutic Goods Administration (TGA), 228 contained
zinc oxide, 363 contained titanium dioxide and 73 contained both (TGA, 2006).
Nano-zinc oxide coatings on clear glass beer bottles prevent UV-degradation of the
contents, while making them appealingly visible to the consumer. Other metal oxides in
common use include titanium and cerium dioxides, while mixed-metal compounds such
as indium-tin oxide (ITO) are currently used in polishing agents for semiconductor
wafers, sunscreen formulas and scratch-resistant coatings for glass (Arabe, 2003).
Carbon-based nanoparticles comprise the second class (Figure 1). This includes
fullerenes, carbon nanotubes (CNTs) and amorphous carbon nanoparticles. The first
fullerene was discovered in 1985, a sixty carbon atom hollow sphere known as the
buckyball was produced by evaporating graphite (Kroto and Walton, 2007). It was
recently revealed that naturally-produced fullerenes have been around for over a billion
years, found in parts per million concentrations in ancient rock formations and believed
to be carried to earth by comets or asteroids (Becker et al., 1996).
a.
b.
c.
Figure 1. Structures of (a) fullerene and (b) single-walled and (c) multi-walled carbon nanotubes
Carbon nanotubes, first produced in 1991, are cylindrical fullerene derivatives that can
be synthesised under controlled conditions to a particular diameter and size, either from
graphite using an arc discharge or laser ablation, or from a carbon-containing gas using
Fate of Manufactured Nanomaterials in the Australian Environment
3
chemical vapour deposition. The multiwalled-products (MWCNTs) are concentric
cylinders up to 10 nm in length and 5–40 nm in diameter. It was later shown that it was
possible to produce single-walled CNTs in the presence of a cobalt-nickel catalyst.
Single-walled CNTs (SWCNTs) have a strength-to-weight ratio that is 460 times that of
steel (Lekas, 2005).
In aqueous systems, carbon nanoparticles aggregate, due to their inherent
hydrophobicity. This limits their use in aqueous and biomedical applications. Much
research has been done to surface modify these nanoparticles to stabilize aqueous
suspensions. Covalent modification, such as the attachment of polyethylene glycol to
SWCNTs (Holzinger et al., 2004), and non-covalent modifications such as the selfassembly of SWCNTs and the phospholipids, lysophosphatidylcholine (Qiao and Ke,
2006), result in very stable carbon nanoparticle suspensions. These modifications have
implications for their use in certain applications as well as repercussions for their fate
and behaviour in the environment.
Figure 2. Schematic representation of mechanisms whereby surfactants help disperse SWCNTs. Top –
SWCNT encapsulated in a cylindrical surfactant micelle, middle – hemi-micellular adsorption of
surfactants on SWCNTs, and bottom – random adsorption of surfactants on SWCNTs (from Ke and Qiao,
2007)
Annual worldwide production of SWCNT is estimated to exceed 1000 tonnes by 2011
(Lekas, 2005). Fullerenes and carbon tubes are produced in large quantities in factories
with capacities as high as 1,500 tonne/y (Frontier Carbon Corporation, www.fcarbon.com; Fullerene International Corporation, www.fullereneinternational.com).
Carbon nanotubes and their derivatives are both used and proposed to be used in
plastics, catalysts, battery and fuel cell electrodes, super-capacitors, water purification
systems, orthopedic implants, cosmetics, conductive coatings, adhesives and
composites, sensors, and components in electronics, aircraft, aerospace and automotive
industries. Increased production results in an increased potential for release to the
environment, either deliberately in discharges or accidentally in spillages, and a greater
possibility of adverse environmental effects. Increased manufacturing volumes also
increase the absolute quantities discharged to the environment as a result of use in
products, and significantly, the quantities that must be disposed of.
4
Fate of Manufactured Nanomaterials in the Australian Environment
The third class comprises nanoparticulate zerovalent metals such as silver, gold and
iron. Nanoparticulate zerovalent iron has been used for some time for the remediation
of waters, particularly groundwaters, as well as sediments and soils (Tratnyek and
Johnson, 2006). It has been used to remove nitrates via reduction, and has most recently
found use in detoxifying organochlorine pesticides and polychlorinated biphenyls
(Zhang et al. 2003). Mobile iron nanoparticles are effective in treatment of dissolved
non-aqueous phase liquids (DNAPL) (Tratnyek and Johnson, 2006).
Bimetallic nanoparticles such as Fe-Pd, Fe-Ni, Fe-Ag, Pd-Ru, etc., have found extensive
use as heterogeneous catalysts (Meyer et al., 2004; Raja et al., 1999).
There is effectively a (voluntary) moratorium on zerovalent iron being used in the UK,
due to unknown potential effects of release of free nanoparticles into the environment
(Royal Society/Royal Academy of Engineering, 2004).
Nanoparticulate silver is one of the most widely used nanomaterials in consumer
products, as indicated in the inventory developed by the Woodrow Wilson International
Centre for Scholars Project on Emerging Nanotechnologies (PEN, 2007a). Applications
are largely based on its bactericidal activity, and include wound dressings, socks and
other textiles, air filters, toothpaste, baby products, vacuum cleaners, and washing
machines. In some cases, the active ingredient is metallic nanoparticulate silver, in
others, ionic silver (Ag+) is electrochemically generated. Ionic silver is not really a
nanoparticle, but is highly particle reactive, so in natural waters is readily adsorbed by
both macroparticles and by colloidal particles such as iron oxyhydroxides or natural
organic matter, and ranges in size from <1 kDa to >0.45 µm (Kramer et al., 2002).
Silver is one element that has useful properties both as a solid and in the dissolved form.
Its antimicrobial activity is most often attributed to the dissolved cation, while it has
entirely different properties as a non-ionic nanoparticle. In both cases, however, the
instability of the monovalent cation and the non-ionic nanoparticle result in extremely
short half-lives of the desired form. This has resulted in research to stabilise silver
nanoparticles to make them useful in biological and other aqueous applications (Doty et
al., 2005). This has created ambiguity in how investigators describe test systems and
manufacturers describe products. For example, it is common for manufacturers to
describe colloidal silver as ‘nanosilver’, rather than metallic silver powder that is
commercially available as nanoparticles.
Colloidal elemental gold has been used for many years, especially in medical
applications as a vector in tumour therapy. Its size varies from 20-160 nm and the
spectral properties change with the classical colour variation from ruby red through
purple to pale blue as size increases (Turkevich et al., 1954). Newer applications of
nanoparticulate gold include its use in electronics in flexible conducting inks or films,
and as catalysts (Haruta et al., 1989).
Fluorescent semiconductor nanocrystals, also known as quantum dots (QDs) form a
fourth class of nanomaterials. Typical materials include CdTe, CdSe/ZnS, CdSe, PbSe
and InP with size ranges from 10 to 50 nm. They usually consist of a semiconductor
core surrounded by a shell, e.g. silica (Sass, 2007). Newer formulations are coming
onto the market that do not have Cd, Pb or Se in the structure and are composed of just
Fate of Manufactured Nanomaterials in the Australian Environment
5
gallium and zinc. Electrons are excited to higher energy levels in the core and the shell,
then fall into the empty spaces left behind. The dot then forms an "exciton" and emits a
particle of light. Changing the size of a QD-based LED makes it emit a different
wavelength of light – producing red, orange, yellow, or green light. The devices are
useful in that they only need about 3 to 4 volts to operate and can run for over 300 hours
without losing any brightness. Their surface is usually functionalised by coatings to
ensure solubility in water.
Synthetic clays represent a large class of nanomaterials with over 9000 tonnes of
nanoclays being produced in 2007 (Electronics.ca, 2007). Both manufactured and
natural clays are starting materials in nanocomposites for use in polymer
nanocomposites, in packaging, paints and cosmetics. They typically range in size from
10 nm to 100 nm. Both negatively charged and positively charged platelets can be
obtained. The former include montmorillinite and hectorite clays, while synthetic
layered double hydroxides (LDHs) of magnesium and aluminium have exchangeable
interlayer anions (Choy et al., 2000; Xu et al., 2006a.b). The anion exchange properties
of these materials allow binding to negatively charged biomolecules between the
hydroxide layers, with hybridisation effectively neutralising the charge and facilitating
penetration into cells, hence their potential as drug delivery agents.
Dendrimers are monodisperse multifunctional polymers that have repeatedly branched
components that form a fifth class of nanomaterials. They are typically spheroid or
globular nanostructures designed to carry molecules encapsulated in their interior void
spaces or attached to their surface (Figure 3). They range in size from around 5 nm for
the simple molecule shown in Figure 3 to five times that and more in larger polymers.
Their synthesis uses repeating procedures to build up their branches from molecular to
the nanoscale (e.g. see Frechet and Tomalia, 2002; Dendritic Technologies Inc,
www.dnanotech.com). These macromolecules can be used for many useful applications
in different fields from biology, material sciences, surface modification, to
enantioselective catalysis.
Figure 3. Typical structure of a dendrimer (first-generation polyphenylene dendrimer reported by Müllen
and coworkers in Chem.-Eur. J., 2002, 3858-3864)
6
Fate of Manufactured Nanomaterials in the Australian Environment
Among the most outstanding applications of dendrimers are the formation of nanotubes,
micro and macrocapsules, nanolatex, coloured glasses, chemical sensors, and modified
electrodes. Some of the uses of dendrimers in biology include DNA transfecting agents,
therapeutic agents for prion diseases, formation of hydrogels, drug delivery, DNA chips,
and ex vivo amplification of human blood cells.
Nanoparticulate emulsions or nanoemulsions are a potential additional class of
nanoparticles, more recently referred to as soft nanoparticles. Emulsions are by
definition dispersions of one immiscible liquid in another and while we are accustomed
to thinking of particles as solid phases, the term really refers to ‘small amounts’ so
could include emulsions. A nanoemulsion has been defined as a type of emulsion in
which the sizes of the particles in the dispersed phase are less than 1000 nm. This
includes particles much larger than the accepted nanoparticle size of < 100 nm, and
typically 20–200 nm, and on that basis, nanoemulsions are excluded from this review.
Nanoemulsions include latex and other formulations used in paints and surface coatings.
They are also used in sunscreens and similar cream formulations, and in medicine for
drug delivery. It is worth noting that the term microemulsion is also used to describe
oil/water emulsions in the nanoemulsion size range and below. The distinction is a fine
one, with microemulsions being thermodynamically stable while nanoemulsions are
kinetically stable (Lawrence and Waresnoicharoen, 2006).
From the list of manufactured nanoparticles and their reported uses, apart from the use
of iron and related bimetallic nanoparticles for water and soil remediation, it appears
that there are few confirmed uses of nanoparticles as agricultural or veterinary
chemicals. So saying, there is potential for use in veterinary medicine for drug delivery
uses and other applications common to human medical uses. One reference highlighted
the use of nanoemulsions for crop applications
(www.nanowerk.com/spotlight/spotid=5305.php).
A distinction has been made by some authors between nanosized particles and
nanosized molecules. The latter include fullerenes and dendrimers. If a molecule
contains segments or has an internal insoluble core, it is considered to be a particle.
Where size is determined by milling, the product will be a nanoparticle. The functional
significance of these separate definitions is not immediately obvious. There will be
differences in fate and toxicity just as there are between different types of non-molecular
nanoparticles.
3.
NANOPARTICLE USAGE IN AUSTRALIA
A voluntary call for information on nanoparticle usage in Australia was recently issued
by the National Industrial Notification and Assessment Scheme (NICNAS). In the
absence of publicly available information, the call was targeted at manufacturers and
importers of nanomaterials or products containing nanomaterials for industrial
(including domestic and cosmetic) use during 2005 and 2006. It was designed as a first
step in understanding the potential for exposure. In addition to issuing an open call in
the Chemical Gazette of February 2006, companies known to be involved with
nanomaterials were directly contacted by NICNAS. The results of the survey were
Fate of Manufactured Nanomaterials in the Australian Environment
7
published on the NICNAS website (www.nicnas.com.au) in January 2007 and are
summarised in Table 2.
Table 2. Usage of nanomaterials in the commercial sector in Australia
Chemical Name
Application
Total usage (tonne/y)
Acrylic latex
Surface coatings
10,000-50,000
Aluminium oxide
Printing
0.05-0.1
Aluminosilicates
Water treatment
10-50
Carbon black pigment
Surface coatings
10-50
Cerium oxide
Catalysts
1-5
Iron oxide
Surface coatings
1-5
Cosmetics
<0.01
Pearl powder
Cosmetics
0.01-0.05
Phthalocyanine
Surface coatings
10-50
Polyurethane resin
Surface coatings
<0.01
Silica dimethylsilylate
Cosmetics
<0.01
Silicon dioxide
Surface coatings
10-50
Water treatment
0.05-0.1
Sodium silicates
Water treatment
0.1-0.5
Surface-treated silicon dioxide
Printing
1-5
Surface-treated aluminium oxide
Printing
0.1-0.5
Surface-treated titanium oxide
Printing
0.5-1
Titanium dioxide
Water treatment
5-10
Domestic products
1-5
Cosmetics
1-5
Surface coatings
5-10
Cosmetics
1-5
Zinc oxide
The interesting finding from this survey, in addition to the expected high usage of
acrylic latex in nanoemulsions, was the fact that CNTs, fullerenes and silver were not
imported or manufactured (as chemicals or in products) at that time, given the high
production volumes projected for 2008-2009 internationally. A second survey is
currently being undertaken.
Despite the fact that many of the literature reports on CNTs may be on proposed uses,
the Woodrow Wilson Project on Emerging Nanotechnologies’ on-line inventory of
nanotechnology-based consumer products (PEN, 2007a) lists 45 carbon-based products
8
Fate of Manufactured Nanomaterials in the Australian Environment
as of February 22, 2008. This is the second most common material after silver (143
references), and followed by zinc oxide (28), titanium dioxide (28), silica (27) and gold
(15). Similarly there is no reference to silver nanoparticle usage in Australia, when we
know it is a component of many consumer products, or to nanoclays. Further local
research is needed to confirm actual usage of CNTs, silver and nanoclays.
4.
CHARACTERISTICS OF ENVIRONMENTAL
NANOPARTICLES
Nanoparticles are characterised by a number of key physical parameters, including size,
shape, surface area, molecular weight (in the case of polymeric particles), and by their
chemical composition. Measurement of these properties is not a trivial exercise as will
be discussed later. The challenge is to determine these properties when the
nanomaterials reach the particular environmental compartments, (atmospheric, aquatic,
terrestrial). Where the measurement technique requires the nanoparticles to be
separated, e.g. electron microscopy, the possibility exists that this will perturb the
natural properties from their form in the environment.
4.1
Manufactured Nanomaterials
The properties of manufactured nanomaterials, as produced, will vary greatly once
released to the environment as interactions occur with other chemicals, or as
transformations such as aggregation and dissolution take place. Such processes can
dramatically affect subsequent biological interactions. Because of this, the concept of
intrinsic toxicity of manufactured nanoparticles is not a useful one, and needs to be
linked with measurements in field or simulated field media. There is a parallel here in
the study of metal speciation in aquatic and terrestrial systems, where the guideline
framework uses a conservative, total dissolved metals as a first cut before a detailed
measurement of a bioavailable fraction. Here the conservative assumption might be to
first base assessments of biological impact on the smallest and potentially more
bioavailable particles in the absence of later more appropriate measurements of actual
size.
The situation becomes more complex because the formulations of manufactured
nanomaterials can often include other additives that can alter the physical behaviour of
nanoparticles in some media, as will be discussed later. Such a concept is not
unfamiliar, for example, in the regulation of pesticide formulations.
4.2
Natural Nanoparticles
It is important to recognise that in both soil, water, and indeed air, compartments there
are a range of natural nanoparticles. In air, there are dusts as well as aerosols
comprising fine particles associated with volatiles emissions from trees and other plants,
or with ‘natural’ events such as bushfires, typically less than 1 µm, but formed by
agglomeration of much smaller particles. Natural clays can be a significant
Fate of Manufactured Nanomaterials in the Australian Environment
9
nanoparticulate component of some soils, as can iron and manganese oxides and other
high molecular weight mineral phases as well as dissolved organic matter in soil pore
waters. In natural waters, as well as in soil pore waters, colloidal particles comprise
clays, iron and manganese oxides and organic matter. We can learn a lot about the
expected behaviour of manufactured nanoparticles from what is already known about
natural nanoparticles
Colloids and macromolecules in natural waters comprise fulvic and humic acids,
fibrillar colloids (exopolymers) that are exudates from algae and other microorganisms
(these are largely polysaccharides and some proteins), and hydrous iron, manganese and
aluminium oxides (Wilkinson and Lead, 2007). Natural colloids fall in the size range
from 1–1000 nm, depending on their degree of aggregation (Buffle and Leppard, 1995a;
Lead and Wilkinson, 2007). Fulvics, humics, and proteins are typically smaller than
tens of nm while polysaccharides and metal oxides are larger, although iron oxide
colloids cover the full size range. Typically these are not present as discrete particles or
compounds, but are heterogeneous mixtures of organic and inorganic species (Lead and
Wilkinson, 2007). Microbial activity in natural waters is a continuous source of
macromolecular material (e.g. polysaccharides) (Buffle and Leppard, 1995b)
Over the wide range of colloid particle sizes, the largest particles have the greatest
percentage mass, but the smaller particles have the greatest number and percentage of
total surface area. Buffle and Leppard (1995a) showed that irrespective of the aquatic
system of interest, the size distribution based on particle number (N) follows Pareto's
Law (i.e. dN/ddp = A dp-b , where A and b are constants with a b value close to 3, and dp,
is particle diameter). The inverse linear relationship between log (particle
number/particle diameter) and log (diameter) means that there are orders of magnitude
more smaller particles than large ones in a water system.
The aggregation of colloids is dependent upon particle size, density, surface charge and
chemical properties (Buffle and Leppard, 1995a; Handy et al., 2008). Aggregation
occurs as a result of particle-particle collisions, involving natural Brownian motion,
different shear velocities in flowing systems and different settling behaviour of different
sized particles.
It has been shown both practically and theoretically, that for a mixture of colloids in
which each size fraction has the same volume, the smallest colloids (<100 nm)
disappear first by aggregation, and the largest by sedimentation, leaving a distribution of
sizes over the range 100 nm-1 µm. (Buffle and Leppard, 1995b).
The interactions between colloids will be governed by their charge and the nature of
their bonding (covalent vs electrostatic). The surface charge of clays at the pH of
natural waters is typically negative over a range of natural pH values. So too is the
charge on most natural organic matter due to ionisable functional groups (e.g. hydroxyl
and carboxylic acid). Iron and aluminium oxyhydroxides have a positive charge below
the pH values at which the surface charge is zero (pH 8-9), however, binding with
natural organic matter typically results in aluminium and iron colloids having a net
negative charge in natural waters (Kretzschemar and Schafer, 2005).
10
Fate of Manufactured Nanomaterials in the Australian Environment
It is not easy to measure surface charge, but it is implied by measurements of the zeta
potential (the potential between the colloid particle surface and solution). As a measure
of the stability of colloidal particles, the zeta potential range between +30 mV and -30
mV is characterised by instability with aqueous dispersions being stable on either side
of that range.
Particles with near neutral charges aggregate rapidly. In natural systems, such
interactions of organic macromolecules and colloidal particles lead to the formation of
loose aggregates or flocs whose structure will be dependent on the relative
concentrations of each in the mix and of the density, shape of the particles and the
flexibility of the macromolecules. Their stability depends on their relative charges and
the nature of the bonding. The aggregates may be stabilised at small sizes that will not
sediment. Larger aggregates form more slowly. It is difficult to predict the behaviour
of such mixtures in terms of rates of reaction and stability, however, it appears that in
low ionic strength solutions, appreciable stability is generally achieved in the size range
100 nm-1 µm as discussed above (Buffle and Leppard, 1995b). Neutrally-buoyant submicron particles can migrate with currents over large distances in fresh waters. The rate
of settling will be controlled by both hydrodynamics and particle size. Deeper, wellmixed waters have reduced settling and larger colloidal aggregates. Once the aggregates
become sufficiently large (>1 µm) they exceed the buoyant mass of colloids and the
newly formed macroparticles will gradually sediment.
Aggregation or particle coagulation can be faster in higher ionic strength water
(seawater) compared to freshwaters, where colloids can be naturally stabilised by
organic macromolecules (Gustafsson and Gschwend, 1997). In estuarine waters,
increasing ionic strength increases screening of the particle charge, resulting in
increased aggregation and coagulation of colloidal particles (Buffle and Leppard,
1995a). For example, the aggregation and precipitation of colloidal iron at salinities
above 15 ‰ (by comparison, seawater has a salinity of 35 ‰) was greater than 75%
complete within 30 minutes, with particles larger than 1.2 µm being formed (Liang and
Morgan, 1990).
A schematic diagram of aggregate formation involving natural colloids is shown in
Figure 4 (from Buffle et al., 1998). This does not consider any living components such
as bacteria and viruses which would add a further layer of complexity.
Natural colloids are frequently in high concentrations in soil pore waters and in natural
water systems, as high as mg/L, so interactions of these particles with manufactured
nanoparticles will be an important fate pathway to consider, and one that is overlooked
in laboratory investigations in synthetic media. The basic behaviour of natural colloids
and macromolecules in soils has been known for decades (Cameron, 1915), and is
governed by the same processes as those in natural waters. High ionic strength (salt
content) in soils will promote flocculation of particles, as will soil pore waters
dominated by calcium and low in sodium (Rengasamy and Olsson, 1991). Many
Australian soils are sodic (sodium rich) (Naidu and Rengasamy, 1993), conditions
which promote dispersion of natural soil colloids when low ionic strength (i.e. low salt)
solution wets the soil (i.e. rainfall or good quality irrigation water) leading to adverse
soil conditions for agriculture e.g. crusting, clogging of soil pores reducing water flow,
Fate of Manufactured Nanomaterials in the Australian Environment
11
reduced aeration (due to poor drainage), etc. These conditions are likely to act similarly
on manufactured nanoparticles, although this needs confirmation.
Figure 4. Major types of aggregates formed in the three-colloidal component system: fulvic compounds
(or aggregated refractory organic material), small points; inorganic colloids, circles; rigid biopolymers,
lines. Both fulvics and polysaccharides can also form gels, which are represented here as gray areas into
which inorganic colloids can be embedded. (From Buffle et al., 1998)
5.
ENVIRONMENTAL SOURCES OF MANUFACTURED
NANOPARTICLES
Unlike many anthropogenically derived nanoparticles, it is reasonable to assume that
there will controls on the release of manufactured nanoparticles that minimise their
release to the environment. The obvious sources that require management are release to
the atmosphere and release via aqueous discharges. These and other potential input
sources are illustrated schematically in Figure 5.
Atmospheric nanoparticles are both a potential risk to the environment and an
occupational health and safety concern for workers engaged in nanomaterial
manufacture. Sources include motor vehicle exhausts and stack emissions from a range
of sources. Eliminating exposure to workplace respirable nanoparticles will be a
priority and is easily addressed through both the use of filters and appropriate protective
clothing and respiratory protection. Motor vehicle and stack emissions are more
problematic. Dealing with fine particle emissions has been an issue for the power
industry for many years, and it is fair to say, has not been adequately eliminated.
12
Fate of Manufactured Nanomaterials in the Australian Environment
Nanoparticle filtration requires nanofilters, which are available, however, appropriate
monitoring will be necessary to ensure their effectiveness.
Atmospheric deposition
Soil application
Effluent discharge
Surface runoff
Road runoff
Groundwater discharge
Accidental spillage
Figure 5. Potential sources of manufactured nanoparticles to the environment
The potential for nanoparticles to end up in aqueous discharges is currently unknown.
Depending on the manufacturing process, there are likely to be both solid and liquid
wastes that may contain nanoparticles. These days, discharges of aqueous wastes are
licensed, although there is unlikely to yet be advice on nanoparticle concentrations.
Similarly uncontrolled disposal of solid chemical wastes is generally not permitted, but
guidance on nanomaterials is most likely absent, so this remains a potential source.
Water treatment plants may have the capability to treat and remove nanomaterials where
discharges are to the sewage system, but as yet there is no information on the ability of
water treatment plants to deal with nanoparticulate contaminants. In particular, anionic
and uncharged nanomaterials could pass through into sewage effluents and not be
retained in sewage biosolids. Several recent studies have indicated a potential for
nanomaterials to interact with bacteria in sewage treatment plants. Choi et al. (2008)
showed that silver nanoparticles were toxic to nitrifying bacteria and that this could
imply detrimental effects on the microorganisms in wastewater treatment. Titanium
dioxide nanoparticles in the presence of ultraviolet light were shown to be toxic to E.
coli, inhibiting the fouling of water treatment membranes (Kwak et al., 2001). Ghafari et
al. (2008) found that SWCNTs caused the protozoan Tetrahymena thermophilia, present
in sewage treatment plants to release excess exudates, which contribute to floc
formation, so they could be used to improve the efficiency of ciliates in wastewater
treatment although effective measures to control and monitor SWCNT release would be
necessary. By contrast, Nyberg et al. (2008) recently indicated little toxicity of
fullerenes in sewage treatment sludge to methanogenic bacteria.
Fate of Manufactured Nanomaterials in the Australian Environment
13
There are instances where nanomaterials are added to aquatic or terrestrial systems for
remediation purposes, e.g. zerovalent iron addition to soils or sediments, and their fate
and impacts will be separately discussed.
The remaining sources are accidental release and release as a consequence of product
usage. The accidental release amounts to spillage of containers, drums, etc., where solid
nanomaterials can end up on land or in water systems. The issue of product usage
requires consideration for each nanomaterial category and for formulations within each
category and this will be discussed in more detail below.
The US Department of Energy has recently published an approach to nanomaterial
environmental safety and health that discusses in some detail the requirements for
nanomaterial transportation and for the management of nanomaterial-bearing waste
streams and nanomaterial spills that minimise the likelihood of releases of
nanomaterials to the environment (USDOE, 2007).
In addressing the risks posed by manufactured nanomaterials, a relevant question is
which nanoparticles have the highest potential for release. Intuitively, these are likely to
be those being produced in the greatest amounts, however, if these productions are from
a large number of widely-dispersed small scale activities, perhaps the risk is less than
from larger facilities with high volume throughputs. Silver fits into the small but
dispersed source category. In particular, the in situ generation of silver nanoparticles in
washing machines will be a highly dispersed source, that may end up in wastewater
treatment plants, and from there may reach the environment although may well be
recovered in the flocculation stages of such plants.
The particular formulation of the nanomaterials is also important for assessing potential
for diffuse releases into the environment. Where nanoparticulates are incorporated into
stable solid-phases, e.g. ZnO nanoparticles in coatings on glass for UV protection, then
the potential for release of the dispersed nanoparticles is low. Where the nanoparticle is
used in a dispersed form (e.g. zerovalent iron for groundwater remediation), then the
potential for movement and effects is much higher.
6.
6.1
FATE OF NANOMATERIALS IN AQUATIC SYSTEMS
Key Pathways
The major physicochemical pathways that govern the fate of nanomaterials in the
aquatic environment are summarised in Figure 6. These comprise aggregation and
subsequent sedimentation, dissolution, adsorption to particulates and other solid
surfaces, binding to natural dissolved organic matter, and stabilisation via surfactants.
Other processes include biological degradation (aerobic and anaerobic), and abiotic
degradation (including hydrolysis and photolysis). Oxidation and reduction may also be
of concern in some environments for specific materials. Concentration in the surface
microlayer of water bodies is a possibility, but unlikely to be a major pathway. The
ultimate fate is likely to involve accumulation and burial in bottom sediments.
14
Fate of Manufactured Nanomaterials in the Australian Environment
In general, the fate of manufactured nanomaterials in aquatic systems has not been that
well studied, however, what information is available, coupled with the extensive
literature on natural colloids in aquatic systems can provide a useful basis for prediction
of nanomaterial fate. The interactions of nanomaterials with natural colloids will play a
critical role in their fate.
Surfactant-stabilised
nanoparticles
Binding to suspended
particles/biota
Binding to NOM
Aggregation
Binding to NOM
and other colloids
and other colloids
Dissolution
Biological degradation,
photolysis, hydrolysis
Mn+
Mn+
Mn+
Sedimentation
Figure 6. Pathways for manufactured metal oxide nanoparticles in natural waters
6.2
Behaviour of Manufactured Nanoparticles
Of the pathways identified in Figure 6, the two most important contributors to the
environmental impacts of manufactured nanomaterials in waters are aggregation and
dissolution.
6.2.1
Aggregation
As shown for natural nanoparticulate colloids (in Section 4.2), the behaviour of
nanoparticles in aqueous systems mimics colloid behaviour. There is a natural
propensity for nanoparticles to grow in size in aqueous solution. Particles that
according to manufacturers’ specifications are nanosized, when suspended in water at
neutral pH, are frequently aggregated (e.g. Figure 7), and the size of these aggregates is
frequently greater than 100 nm, the upper boundary of the nanoparticle size range.
In the case of nanoparticles with a surface charge, screening of the surface charge by
electrolyte ions, e.g. in seawater, overcomes the electrostatic forces and allows
aggregation, as for natural colloids. Steric stabilisation of nanoparticles against
Fate of Manufactured Nanomaterials in the Australian Environment
15
aggregation can occur through surface modification by surfactants or bulky polymeric
additives.
Interactions of nanomaterials with natural colloids (organic macromolecules, inorganic
colloids or heterogeneous aggregates) will also occur in the same manner as discussed
in Section 4.2 (Saleh et al. 2008).
The rates at which manufactured nanoparticles aggregate is particularly important, since
the slower the aggregation the greater the potential for interaction with biota.
Unfortunately this has been poorly studied, although there are data available for natural
colloidal nanomaterials.
The terms aggregate and agglomerate have distinct meanings in particle science, but are
frequently confused. As discussed by Nichols et al. (2002), agglomerates are generally
considered to be an assemblage of particles that are rigidly bound by fusion sintering or
growth, while aggregates are loosely bound particles that are readily dispersed. The
word clump is also used but they proposed replacement of ‘clumps’ with ‘agglomerates’
that may be hard (not readily dispersed) or soft (readily dispersed).
A summary of the available data on particle aggregation is presented in Table 3.
Brant et al. (2005) reported that n-C60 fullerenes (i.e. nanoscale suspended aggregates
known as fullerene water suspensions), showed a strong tendency to aggregate in weak
electrolyte solutions greater than 0.001 M ionic strength. Below these concentrations,
aggregates were stable for over 15 weeks (Lyon et al., 2006). The same effect of ionic
strength on natural colloid aggregation was noted earlier. The n-C60 aggregates
eventually settle out of suspension, sorb to particles or become otherwise immobilised
on surfaces.
Table 3. Aggregation data for manufactured nanomaterials in water (adapted from Boxall et al., 2007)
Nanomaterial
Water type
Aggregate size
range, nm
Comments
References
Fullerenes
Freshwater
culture
medium
25-500 (mean 75)
This is with a THFbased preparation.
Smaller mean size
using sonication
Lyon et al., 2006
TiO2
Freshwater
177-810 (mean
330)
From an initial size
of 66 nm
Adams et al., 2006
SiO2
Freshwater
135-510 (mean
205)
From an initial size
of 14 nm
Adams et al., 2006
ZnO
Freshwater
420-640 (mean
480)
From an initial size
of 67 nm
Adams et al., 2006
Zerovalent
iron
Freshwater
1000
Rapidly aggregate
Mondal et al.,
2004; Schrick et
al., 2004
16
Fate of Manufactured Nanomaterials in the Australian Environment
1 µm
100 nm
Figure 7. Electron micrographs illustrating aggregation of zinc oxide nanoparticles from dispersion of a
ZnO nanopowder (nominally 30 nm) in a freshwater algal medium, pH 7.5
Zerovalent iron nanoparticles in water grow rapidly to micron sizes or more, and
quickly lose reactivity, rapidly settling out of solution (Phenrat et al., 2004).
In general, the effect of basic water chemistry (pH, redox potential, hardness, salinity)
on nanoparticle stability has been poorly studied. Lead et al. (2007) showed, for
example, that aggregation of gold (and iron oxide) nanoparticles was minimised at low
pH. While such studies assist in understanding aggregation behaviour, they are of little
value in predicting the behaviour in natural water systems where pH variation is limited.
There have been attempts to develop predictive models for aggregation behaviour
(Mackey et al., 2006), but these are as yet untested, and given the complexity of natural
waters, their applicability may be problematic.
A number of papers have documented oxidation/reduction reactions of fullerenes, and
the potential for oxidation (hydroxylation) mediated by fungal enzymes has been
suggested (Wiesner et al., 2006). No such biotransformations have, however, yet been
observed.
6.2.2
Nanoparticle Solubility
With respect to solubility, the Gibbs-Thompson effect predicts that nanoparticles with a
smaller radius of curvature are energetically unfavourable and subject to preferential
dissolution, and have a higher equilibrium solubility than macroparticles (Figure 8)
(Borm et al., 2006). This solubility can exceed saturation conditions in some instances,
leading to growth and precipitation of particles in a phenomenon known as Ostwald
ripening, where with time, the rapid initial dilution and supersaturation solubility is
reduced by the growth of larger particles with lower solubility. The overall process is
one of destabilisation of nanoparticles in solution.
Fate of Manufactured Nanomaterials in the Australian Environment
17
These phenomena raise questions about the overall stability of nanoparticles in aquatic
environments and highlight the need for measurements of both particle size and
solubility to reliably assess the fate of nanomaterials.
Most metal-based nanoparticles are hydrophilic and have a finite but often low
solubility. In many cases, this is not measured, but since the soluble ionic metal fraction
is the most toxic to aquatic biota, it is desirable that the extent of this solubility be
determined. As a case in point, Franklin et al. (2007), investigating the biological
impacts of zinc oxide nanoparticles, found that despite a common belief that zinc oxide
was ‘insoluble’, nanoparticulate ZnO rapidly dissolved to the extent of 6 mg/L of
dissolved (dialyzable) zinc within 6 h and 16 mg/L in 72 h in a buffered pH 7.5 algal
medium. This was a concentration well in excess of the 5 mg Zn/L that would be toxic
to most aquatic biota. By contrast, in similar experiments with nanoparticulate cerium
oxide, a very low solubility (ng/L) was observed, and so the effects of nanoparticle
versus macroparticle toxicity could be readily investigated (Franklin, unpublished
results). Greater toxicity to algae was observed for nanoparticulate CeO2 compared to its
macroparticulate equivalent.
Figure 8. Illustration of the solubility of amorphous silica as a function of radius of curvature (adapted
from Bjorn et al., 2006)
The toxicity of a range of metal nanoparticles to a range of aquatic organisms was
investigated by Griffitt et al. (2008). Toxicity was observed for silver and copper with
48-h LC50s to Daphnia pulex being 40 and 60 µg/L respectively. Here the role of
18
Fate of Manufactured Nanomaterials in the Australian Environment
dissolution was demonstrated to be minor, however, solubility played a major role in the
toxicity of nickel nanoparticles.
Semiconductor quantum dots based on cadmium selenide have been shown to release
ionic cadmium as a result of selenide oxidation (Derfus et al., 2004). Solutions of 250
mg/L yielded as high as 80 mg Cd/L. The concentration of cadmium directly correlated
with cytotoxic effects to primary hepatocytes isolated from rats and grown in vitro.
CdSe/ZnS nanocrystals released a factor of 10 less cadmium, however only
polyethylene-silane coatings were effective in preventing release (Kirchner et al., 2005).
In the case of environmental release, such concentrations of cadmium would readily
exceed water quality guidelines (ANZECC/ARMCANZ, 2000) with severe
consequences for ecosystem health.
Carbon-based nanoparticles are typically lipophilic and are virtually insoluble in natural
waters. The solubility of fullerene has been calculated as 10-18 mol/L (Abraham et al.,
2000). The lipophilicity will vary with substitution on the basic fullerene or nanotube
formulations, and derivatives have been prepared with appreciable water solubility.
Sayes et al. (2004) showed that cytotoxicity to human liver carcinoma cells was
inversely related to the solubility of fullerene derivatives, largely as a consequence of
the reduced ability to generate oxygen free radicals that are the cause of cytotoxic
effects via lipid peroxidation.
It is important to recognise that the term ‘solubility’ has been loosely used by some
authors, especially in relation to carbon-based nanomaterials, often meaning forming
stabilised suspensions as distinct from truly dissolving as was the case with metal
oxides for example.
6.2.3
Role of Nanomaterial Formulations and Impurities
In many instances, the formulations of nanomaterials include additives (e.g. surfactants),
which are added to modify the surface properties, and to minimise aggregation. These
formulations may also result in different solubility characteristics. Carbon nanotubes
are extremely hydrophobic and subject to high van der Vaal’s forces along the length
axis, with a tendency to aggregate. To disperse CNTs in aqueous solution, a range of
chemical additives have been used including surfactants (sodium dodecylsulfate,
sodium dodecylbenzene sulfonate, Triton X-100) and polymers, acting either sterically
or electrostatically (Brant et al., 2005). The effect of these dispersant additives is
usually to sterically stabilise the nanomaterials, by physically hindering their
aggregation (Handy et al., 2008; Saleh et al., 2008). Terashima and Nagao (2007)
showed that the surfactant Triton X-100 and natural humic substances enhance the
solubility of C60 nanoparticles by 8-540 times, while also decreasing the rate of
aggregation (Chen and Elimelech, 2007).
In many cases the effect of additives on solubility and aggregation in commercial
nanomaterial formulations is unknown. In the case of zinc oxide formulations, for
example, a relevant concern might be whether the equilibrium water solubility of
nanoparticulate zinc oxide in sunscreens is different to that seen for the raw
Fate of Manufactured Nanomaterials in the Australian Environment
19
nanoparticles by Franklin et al. (2007). Such questions have important implications for
risk assessment in aquatic systems.
Nanomaterials often contain impurities, for example, carbon nanotubes have been
reported to contain metal catalyst impurities (Haddon et al., 2004). Plata et al. (2008)
showed that metal and carbonaceous impurities could account for up to 70% of the
weight of the SWCNT formulations, with nickel up to 22%, yttrium 6%, cobalt 2-10%,
iron 0.5% and molybdenum 0.7%, together with traces of copper, lead and chromium.
Amorphous carbon could be as high as 45%, while polycyclic aromatic hydrocarbons
(particularly naphthalene) were up to 60 µg/g in arc-produced nanotubes and even
higher in those prepared by chemical vapour deposition. These impurities can affect the
surface charge, reactivity, transport and ecotoxicology of the SWCNTs.
The presence of impurities was found to be responsible for oxidative stress damage to
rat epithelial cells (Pulskamp et al., 2007). Similarly, the presence of tetrahydrofuran
(THF) residues was shown to be responsible for observed toxicity of n-C60 fullerenes to
large mouth bass (Brant et al., 2005). While these impurities are not expected to
significantly affect nanoparticle fate, they raise interesting questions with respect to the
effects on the environment. It could be argued that THF-containing n-C60 and metalfree SWCNTs are both unnatural forms, and therefore are not environmentally relevant,
but if these are present in the manufactured products then their behaviour is a valid
concern.
6.2.4
Fate in Natural Water Systems
While there are data from laboratory studies on the behaviour of selected nanomaterials
in water, the behaviour is likely to differ in natural waters, where there is a possibility of
interaction with natural colloids including dissolved (and particulate) organic matter
(NOM). The importance of colloids cannot be underestimated. In freshwaters for
example, colloidal organic matter concentrations lie in the range 1-10 mg/L compared to
the concentrations that have been predicted for manufactured nanoparticles of 1-100
µg/L, which is at least several orders of magnitude lower (Boxall et al., 2007).
A recent study by Hyung et al. (2007) showed that the addition of standard Suwannee
River humic acid greatly enhanced the dispersion of multi-walled carbon nanotubes in
Milli-Q water, and that the same effects were also seen in suspensions in Suwannee
River water samples. The dispersion was greater than that observed with sodium
dodecylsulfate. The exact mechanism of the enhanced dispersion is likely to again
involve both steric and electrostatic components, as was seen for natural colloids.
Similar stabilisation of iron oxide nanoparticles by humic acids has also been
demonstrated (Tipping and Higgins, 1982; Baalousha et al., 2008).
By contrast, it has been suggested that natural fibrillar colloids are likely to increase
aggregation because of different binding characteristics, compared to the charge
stabilisation mechanism of humic substances (Buffle et al., 1998).
The findings to date suggest that in natural water systems, nanoparticles may have a
greater stability than in synthetic (NOM-free) waters, particularly in estuarine and
20
Fate of Manufactured Nanomaterials in the Australian Environment
marine waters of higher ionic strength. In waters with a high suspended sediment load,
however, association of nanoparticles is likely to provide an effective removal
mechanism that could enhance transport to and accumulation in bottom sediments.
Given these many uncertainties, site-specific fate studies are recommended that use
actual nanomaterial formulations in a variety of natural waters (fresh and estuarine).
Where nanoparticles are released with wastewaters, it has been suggested that the
presence of household or industrial detergents would result in the disaggregation of
nanoparticles (Fernandes et al., 2006). In a related study of cerium oxide nanoparticles
in a model wastewater treatment system, Limbach et al. (2008) found that a small but
significant fraction (6%) avoided aggregation and was released in the effluent (at 2-5
mg/L concentrations) largely as a result of stabilisation in the presence of protein
breakdown products and surfactants in the wastewater changing the zeta potential.
These examples highlight the impact that surface modifications from wastewater
components can have on nanoparticle fate.
Sinks and issues of non-steady state thermodynamics influence the fate of nanoparticles.
Adsorption of molecules or ions on nanoparticles can catalyse or promote dissolution,
e.g. via chelating agents. This is a dynamic process.
A recent paper by Benn and Westerhoff (2008) revealed some interesting findings on
the fate of nanoparticle silver released into water from commercially available sock
fabrics. Repeated washings released most of the silver, with 70-90% in an ionic form,
and the remainder as large nanoparticles (100-200 nm). In a simulated water treatment
process they showed that all of the silver was removable to the sludge, raising concerns
about the impacts of application of sludge to land.
The behaviour of emulsions in natural waters has been poorly studied. In a report on
acrylic latex, NICNAS (2000) noted that ‘the fate of the aqueous residues released to the
sewer system is less predictable as the notified polymer may remain in the aqueous
phase as an emulsion at low concentrations’. In addition, ‘all solid residues will remain
associated with the soil and sediment due to the high molecular weight and the stability
of the cured paint matrix’.
6.2.5
Nanoparticles as Vectors for Contaminant Transport
While so far we have considered manufactured nanoparticles as potential sources of
toxic effects in the environment, as noted earlier with respect to colloids, nanoparticles
are excellent binding sites for other soluble contaminants and therefore have the
potential to act as vectors for the delivery of these contaminants. Again, this will be a
function of surface properties of the particular nanomaterial formulations.
A recent publication by Hu et al. (2008) showed that aqueous suspensions of fullerene
were able to effectively sorb polycyclic aromatic hydrocarbons (PAHs), a process that
was further enhanced by the addition of humic acids. This predictable behaviour
indicates that nanomaterials can affect the fate of hydrophobic organic contaminants in
natural waters.
Fate of Manufactured Nanomaterials in the Australian Environment
21
In soils, groundwaters, rivers and lakes, natural colloids have been shown to play an
important role in trace metal retention and transport (Kretzschmer and Schafer, 2005).
Similar binding capacities exist for manufactured nanoparticles. Secondary toxicity
effects from these adsorbed contaminants will need to be considered in any toxicity
studies of nanoparticles.
6.3
Fate of Manufactured Nanomaterials in Terrestrial
Systems
There is currently very little information available with which to assess the
environmental risk of manufactured nanoparticles to terrestrial ecosystems. The key
physico-chemical properties of nanoparticles described above are also likely to play a
major role in the fate, transformation, and environmental effects in soils. Soils differ
from fresh and marine waters in that the solid phase provides a large and reactive “sink”
for nanoparticles, so that the applied dose may overestimate the actual dose to soil biota.
One of the key hurdles in examining nanoparticles in terrestrial systems is the detection
of the manufactured nanoparticles in the presence of natural nanoparticles, which are
ubiquitous in soil.
6.3.1
Key Pathways
A number of key processes are likely to affect the fate and bioavailability of
nanoparticles in the soil environment (Figure 9).
Nanoparticles have high surface reactivity and, depending on surface charge and
coatings, their adhesion to reactive soil surfaces may be strong –“partition coefficients”
for nanoparticulate contaminants in soil have yet to be published. Data from transport
studies of soil colloids however indicate that surface coatings on the nanoparticles are
important determinants of mobility and may enhance transport (Kretzschmar et al.,
1995; Seaman and Bertsch, 2000; Saleh et al. 2008), and this has also been found for
nanoparticles used in groundwater remediation (Hydutsky et al,. 2007). As yet, there
are few data on transport of nanoparticles through soils, and hence characterisation of
nanoparticle mobility and associated potential bioavailability remains to be elucidated.
Recent studies examined the transport of eight nanoparticles (fullerol (C60-OHm),
SWCNTs, silica (57 nm), alumoxane, silica (135 nm), n-C60, anatase and ferroxane)
through spherical glass beads and found the attachment efficiencies to fall in the order
as listed (Lecoanet et al., 2004). Another recent sand column study demonstrated the
importance of surface coatings in the transport of zerovalent iron nanoparticles (Saleh et
al., 2008). Similar studies in soils are now required.
22
Fate of Manufactured Nanomaterials in the Australian Environment
1.
2.
3.
4.
5.
6.
7.
Dissolution
Sorption/aggregation
Plant bioaccumulation
Invertebrate accumulation and toxicity
Microbial toxicity
Direct particle uptake/toxicity
Particle migration
MNPs
6
2
3
1
7
4
Dissolved
pool
5
Figure 9. Key processes in soil relating to transformation and potential risk from manufactured
nanoparticulate particles
6.3.2
Behaviour of Natural Colloids in Soils
Naturally-present colloids and macromolecules in soils are similar to those found in
freshwater systems, and nanoparticulate and microparticulate clays, organic mater, iron
oxides and other minerals play an important role in biogeochemical processes. Soil
colloids have been studied for decades in relation to their influence on soil development
(pedogenesis), and their effect on soil structural behaviour (dispersion and crusting)
(Cameron, 1915). Dispersion of soil is a key process affecting the quality of surface
waters in Australia, and studies have examined the factors responsible for colloid
generation and transport in soil systems (Noack et al., 2000; Siepmann et al., 2004;
Kaplan et al., 1996; Seaman et al., 1997).
A large body of literature exists on the aggregation/dispersion behaviour of soil colloids
in relation to soil physical and chemical properties (for a review see Rengasamy and
Olsson, 1991). Aggregation of colloids in soil is a function of surface charge, ionic
strength, particle size and chemical composition of the soil pore water and exchangeable
ions held on the surface of colloids. Systems dominated by sodium and with low ionic
strengths are likely to have dispersion of colloids, while those dominated by calcium
and high ionic strengths will tend to aggregate. Recent evidence confirms that
manufactured nanoparticles behave similarly to natural colloids (Saleh et al. 2008;
Wang et al. 2008). High water flow through soils will tend to mobilise colloids, while
Fate of Manufactured Nanomaterials in the Australian Environment
23
slow water flow will tend to allow interaction and binding of colloids with soil minerals
and organic matter.
6.3.3
Behaviour of Manufactured Nanoparticles in Soils
Of the pathways identified in Figure 9, the most important properties that will control
nanoparticles fate in soils are likely to be dissolution, aggregation and partitioning
between solution and solid phases.
Nanoparticle Solubility
Dissolution of nanoparticles in aqueous media has already been covered in Section 6.2.2
above. A key difference in soils is the large surface area and exchange capacity for
cations and anions that can promote dissolution of compounds through acting as a sink
for dissolution products, and providing protons to enhance dissolution of compounds
with a pH-dependent solubility. To date, no studies have examined the rate or extent of
dissolution of nanoparticulate materials in soils in relation to their bulk counterparts.
Aggregation
There are virtually no studies which have examined this topic for manufactured
nanoparticles in soils, but inferences from the behaviour of natural colloids can be made
(Section 4.2 above). Aggregation behaviour of nanoparticles in aquatic systems has
been covered in Section 4.2, and the same processes would be active in soils, except we
can speculate that the aggregation of nanoparticles in soil may be greater due to the
higher ionic strength of soil pore waters compared to most surface water systems
(streams and dams). Aggregation in soils also leads to particle entrapment in pores
through which the dispersed nanoparticles could have passed, thus restricting mobility
(Wang et al. 2008).
Partitioning
There are virtually no studies which have examined this topic. We can speculate that
the high surface area and charge of many hydrophilic manufactured nanoparticles will
cause a strong binding to the predominantly negatively charged surfaces of soil minerals
and organic matter (Li et al. 2008), depending on the nature of the charge. Net
positively charged particles will be retained strongly, while those with net negative
charge will be highly mobile in most soils (Saleh et al., 2008).
Where nanoparticles are hydrophobic, retention to organic matter surfaces in soil may
inhibit mobility and availability to organisms.
The characteristics of the surface “functional” coatings used in nanoparticle
manufacture may be very important in explaining (and predicting) fate in soil, as it is
these surfaces that will interact with minerals and organic matter surfaces in soil.
Given that contaminant partitioning (Kd, Koc or Kow) is a key property used in risk
assessments for a wide range of inorganic and organic contaminants in terrestrial
systems, this characteristic is a key property requiring evaluation.
24
Fate of Manufactured Nanomaterials in the Australian Environment
7.
ECOLOGICAL RISK ASSESSMENT OF MANUFACTURED
NANOPARTICLES
Regulators worldwide are seeking to undertake ecological risk assessments of
manufactured nanoparticles to determine the significance of any impacts associated with
their manufacture and use. The ecological risk assessment framework for contaminants
in the environment, developed by the USEPA and adopted in Australia, has the
following components:

Problem formulation

Exposure assessment – Chemical assessment taking into account contaminant
fate (predicted environmental concentrations – PECs)

Effects assessment – Measurement of toxicity, bioaccumulation, effects on
ecology (predicted no effect concentrations – PNECs)

Risk characterisation (PEC/PNEC)
As discussed by Owen and Handy (2007), the issue of problem formulation is a critical
one. The initial anxiety that nanomaterials might represent the current equivalent of
genetically modified foods in terms of its environmental danger (Dowling, 2005)
appears to have now passed. These concerns were heightened by the findings that
fullerenes were capable of crossing the blood-brain barrier in fish (Oberdortser, 2003),
which has since been shown to be an experimental artefact (Brant et al., 2005).
Nevertheless there are a number of basic concerns that need addressing, starting with the
basic issue of whether nanosized materials pose a greater hazard to biota than the
equivalent macrosized materials. While there is good evidence for altered behaviour
with smaller size, only a handful of studies have demonstrated that this translates into
greater toxicity. The risk assessment needs to show connectivity between the source,
the pathway, and the receptor. In most instances in water and soil ecosystems, the
evidence of this connectivity has been indirect or absent.
For industrial chemicals, a manual providing guidance on ecological risk assessment
was recently released by the Department of the Environment and Water Resources (now
Department of the Environment, Water, Heritage and the Arts, DEWHA) (DEW, 2007).
This manual specifically discussed data requirements, data evaluation, environmental
exposure assessment, environmental effects assessment, assessment of persistent,
bioaccumulative and toxic substances, and risk characterisation and management.
Data requirements include melting point, specific gravity, vapour pressure, water
solubility, hydrolysis as a function of pH, octanol/water partition coefficient, adsorption
behaviour in soils, acid dissociation constant, and environmental fate data, especially on
biodegradation and bioaccumulation. For effects assessment, toxicity tests must be
undertaken using a fish acute test, Daphnia immobilisation and reproduction tests, an
algal growth inhibition test, and measures of biodegradability and bioaccumulation. As
the following pages will show, the majority of these requirements could not currently be
met for manufactured nanoparticles. This means that the determination of both PECs
Fate of Manufactured Nanomaterials in the Australian Environment
25
and PNECs will not be possible as a prerequisite to assessing the potential
environmental hazard of manufactured nanoparticles in soil, water and sediment
compartments. The current state of knowledge in these areas is reviewed in the
following pages of this report.
7.1
Polymeric Nanoparticles as a Separate Class
Some authors have drawn the distinction between polymers such as dendrimers,
fullerenes and carbon nanotubes whose size is determined by their molecular weight and
other particles where size is a function of their degree of aggregation. There is potential
for the size of both nanoparticle types to affect their interactions with aquatic biota, but
to date few studies have investigated this. We see no reason to consider polymers as
warranting separate regulatory consideration from other nanoparticle types.
It is clear that assessment of hazard of either type on the basis of intrinsic chemical
properties is inappropriate and there must be some consideration of size, be it molecular
weight or other measures of size. The premise that nanoparticulate size fractions are
more toxic than larger size fractions needs to be tested for all nanoparticle classes with
respect to the natural environment into which they are released.
8.
8.1
EXPOSURE ASSESSMENT
What to Measure
An exposure assessment seeks to determine the concentrations and bioavailable forms
of a contaminant in the environment that, with a consideration of fate and exposure
duration, can be linked to effects on target organisms. It will be important therefore that
measurements reflect the concentrations and physical and chemical properties of the
nanoparticles in the field that are truly representative of exposure. In assessing
industrial chemicals, the DEW manual (DEW, 2007) lists fate, partitioning behaviour,
and persistence as important parameters. These need to be combined with concentration
data in estimating likely exposure.
For nanomaterials, since it has been demonstrated that size is a critical parameter, any
measurement of concentration must be accompanied by data on the distribution of
particle sizes in the test water taking into account any aggregation that might occur
within the life cycles of the test organisms.
The requirement and the current status of methods for nanoparticle analysis and
characterisation have been well summarised in recent reviews by Hassellov et al. (2008)
and Tiede et al. (2008). Particle size measured as a diameter was not adequate when
particles were other than spherical, and other measures including aspect ratio (ratio of
their longer dimension to their shorter dimension) were also of value. They believed
that in addition to particle size distributions, measures of surface area were also
important, but not always reported. Nanoparticle net surface charge was also seen as an
26
Fate of Manufactured Nanomaterials in the Australian Environment
important measure of the extent to which their dispersion is stabilised by electrostatic
repulsive forces.
An interesting issue is the extent to which nanoparticle size distributions reach steady
state, and whether this state is maintained throughout the duration of an experiment, e.g.
for chronic toxicity testing. Frederici et al. (2007) noted a change in distribution when
studying the effects of nanoparticulate titanium dioxide on rainbow trout. Hassellov et
al. (2008) recommended that monitoring be undertaken over the duration of any studies
to detect this, or any changes due to other reaction and/or degradation pathways.
8.2
Methods for Measurement of Nanoparticles
The measurement of nanoparticles in environmental media poses particular challenges.
Measurements are required of concentrations and size, and possibly also of surface area
and charge. Where possible, measurements should be of the state of the nanoparticles in
the particular medium (soil, sediment, water), rather than an assumption based on
dilution of a starting material. Even the measurement of concentration poses issues, as
we are concerned with the bioavailable concentration of nanoparticles. In many cases,
what is measured is a surrogate for this, e.g. total zinc concentration rather than
nanoparticulate zinc oxide. In the case of n-C60, UV absorbance was used to measure
concentrations (Oberdorster et al., 2004), while for carbon nanotubes, light scattering
techniques were used to correlate with concentration (Smith et al., 2007).
The bioavailable fraction can however include a dissolved, soluble fraction rather than a
nanoparticulate fraction, so ideally some measurement that discriminates this fraction is
required. Standard 0.45-µm membrane filtration will not retain most nanoparticles, so a
separation technique is required. Ultracentrifugation, size-based chromatographic
separations, ultrafiltration and dialysis are all appropriate, although the last two are
probably the preferred methods of separation. In soils, there is the additional
complication that any nanoparticulate material that dissolves will interact with the soil
solid phase, and some assessment of this pool may also be required to assess
bioavailability in addition to characterisation of the material in soil pore water.
For measuring particle size distributions, electron microscopy (EM) and dynamic light
scattering (DLS) are the most commonly used techniques. Both have advantages and
disadvantages (Bootz et al., 2004). EM gives the most direct information on the size
distribution and shapes of particles, however, there is concern about artefacts introduced
by the sample preparation step. With DLS, the presence of small amounts of large
aggregates can affect the distribution of a main component of a smaller size, with results
being misleading where the samples have a broad size distribution. More detailed
information on specific surface area, surface charge and zeta potential can be obtained
by a variety of techniques, but these are research techniques that are not likely to
contribute to routine risk assessment of nanomaterials in the near term.
For studies of nanoparticles in situ, field flow fractionation (FFF) has been advocated
(Hasselov et al., 2008; Tieded et al., 2008), in particular a variation called flow field
flow fractionation (FlFFF). Basically the technique uses two right-angled flow streams
Fate of Manufactured Nanomaterials in the Australian Environment
27
to partition particles on the basis of their diameter (Giddings, 2003). For metalcontaining nanoparticles, the metal concentrations in the separated fractions can be
analysed by inductively coupled plasma mass spectrometry (ICPMS). Stolpe et al.
(2005) have described the application of high resolution ICPMS coupled to FlFFF to
study metals in (natural) nanoparticulate colloids. The FFF technique has been well
established, but is not that easily mastered, and is in use in only a handful of laboratories
worldwide. The universal interest in nanomaterials might lead to a wider acceptance.
Single particle ICPMS analysis has recently been applied to the detection of gold
colloids in water (Degueldre et al., 2006). The use of new generation ICPMS
approaches for analysing individual nanoparticles show considerable promise (Stolpe et
al., 2005), but it may be some time before they can be applied to routine environmental
monitoring of manufactured nanoparticles.
Similarly the use of novel techniques such as liquid chromatography coupled to nuclear
magnetic resonance spectrometry may hold promise for the analysis of carbon-based
nanoparticles.
The analysis of manufactured nanoparticles in natural systems can be complicated by
the background of natural colloids. Where nanoparticle shape is distinctive, e.g. CNTs,
this may not be such an issue, but for others, techniques such as DLS and FFF will not
be able to discriminate between nanoparticles and natural colloids unless linked to some
nanoparticle element-specific analyses such as ICPMS. The solution is to use single
particle confirmatory analyses, such as EM, with energy-dispersive x-ray fluorescence
(EDX).
To date there have been few measurements of manufactured nanoparticles in natural
waters or soils because of the extreme difficulty in detecting environmental
concentrations. Details of techniques applied to environmental nanoparticles in aquatic
systems have been discussed by Wiggington et al. (2007).
The area of nanoparticle measurement is one that is being pursued internationally by a
number of agencies. In particular, there is a need to develop standard methods of
analysis, including methods for sample preparation that can be used to characterise
nanoparticles. As part of this exercise, the development of standard reference materials
that can be used for method quality assurance and quality control will be essential. In
Australia, the National Measurement Institute (NMI) has an active program in this area.
8.2.1
Relevance of OECD Test Guidelines
The assessment of the environmental fate of chemicals and polymers currently relies on
a few critical tests recommended by the Organisation for Economic Cooperation and
Development (OECD) (OECD, 2007), including those for water solubility,
adsorption/desorption, water/oil partition coefficient, hydrolysis, surface tension and fat
solubility. The applicability of each of these tests to nanomaterials is generally
inappropriate, and the methods will need to be considerably altered to adequately cater
for nanomaterials.
28
Fate of Manufactured Nanomaterials in the Australian Environment
The test for water solubility (No. 105) (OECD, 2007) uses either a microcolumn
separation or a flask dissolution. Since the tests were not designed for use with
colloidal or nanosized particles, the separation of these from the ‘soluble’ fraction will
be critical. The test method indicates that the presence of colloids in the microcolumn
effluent invalidates the test. In studies of zinc oxide solubility, Franklin et al. (2008)
used dialysis to separate soluble zinc. Such procedures will be required as a finish to
the OECD test. The same applies to Test No. 120, for the solution/extraction behaviour
of polymers in water.
The tests looking at adsorption/desorption onto soils need to be relevant to the likely
environmental concentrations. Test No. 106 uses a soluble chemical fraction, however,
for a nanomaterial suspension, this would not be appropriate. Test No. 121 determines
the adsorbed fraction by HPLC. Determining whether the nanoparticles are retained by
filtration rather than adsorption will be problematic.
The octanol/water partitioning tests (Nos 107 and 123) are designed to measure the
equilibrium partitioning of a ‘dissolved’ substance between the two solvents, as distinct
from ‘solubility in octanol’ which is what will be obtained using nanoparticles. Even if
it were meaningful, the physical application of this test is likely to be seriously impaired
by clumping of nanoparticles at the solvent interface.
The hydrolysis test (No. 111) looks at hydrolysis in the range pH 4-9. With
nanomaterials, the result would test both dissolution and hydrolysis as a function of pH.
Surface tension (No. 115) is inappropriate for an insoluble chemical in nanoparticulate
form, however, the assessment of fat solubility (No 116) is a potentially useful measure
in relation to biological uptake.
OECD has an active interest in nanomaterials, and has a working group considering
appropriate test methods (see Section 10.1.3), including those for toxicity testing.
8.3
Modelling Exposure
Existing models of exposure for soluble contaminants have little applicability to
nanoparticles. As already discussed, there have been preliminary approaches to
predictive modelling of the suspension stability and kinetics of aggregation of
nanoparticles, however their applicability to real systems is, as yet, untested (Mackay et
al., 2006).
In an attempt to model likely concentrations of manufactured nanoparticles that might
be found in the environment, Boxall et al. (2007) used a series of simple algorithms to
predict the likely concentrations that might be found in soils and waters. For waters,
they considered five routes of entry:
(i)
the direct entry of manufactured nanoparticles into water bodies from
bioremediation;
(ii)
inputs from spray drift following use of agrochemicals;
Fate of Manufactured Nanomaterials in the Australian Environment
29
(iii)
runoff from contaminated soils;
(iv)
aerial deposition; and
(v)
emissions from wastewater treatment plants.
For soils, routes comprised:
(i)
the application of remediation technologies;
(ii)
the application of plant protection products;
(iii)
the excretion of nanomedicines used in veterinary products;
(iv)
aerial deposition; and
(v)
the application of sewage sludge as a fertiliser.
They focussed mainly on cosmetics, personal care products and paint, and the
nanoparticle concentrations that they contained (based on limited European data). Three
hypothetical scenarios were modelled, where 10, 50 and 100% of a product type
contained the manufactured nanoparticle. Predicted concentrations for the 10% scenario
are shown in Table 4. Despite all of the uncertainties, the concentrations can be
compared to the toxic concentrations where these are known, to see whether these are in
the same range or not.
Table 5 shows the comparison of exposure data with known toxicity data, indicating
that the predicted environmental concentrations are orders of magnitude below those
known to have environmental effects on aquatic biota (as will be elaborated on later).
This scenario naturally does not take into account all possible sources, or accidental
releases. The results nevertheless give regulatory agencies some reassurance, especially
since the assumptions in estimations are conservative.
The challenge for modellers in the derivation of appropriate PECs is to be able to obtain
reliable estimates of the mass flow of nanomaterials to different compartments of the
environment. A good example of a life cycle assessment approach to this is shown in
Figure 10 (from Blaser et al., 2008). This example has been used for silver derived
from nanoparticulate biocidal plastics and textiles, but the approach has generic
application. In deducing mass flows, estimates of total product usage and estimated (or
measured) release rates must be obtained. These data are then related to the time of
exposure. Knowledge of the behaviour of silver in the aquatic environment (colloidal
forms, attachment to particles, etc.) is used in coupled river fate models to predict
sediment/water partitioning during treatment and in the aquatic environment.
30
Fate of Manufactured Nanomaterials in the Australian Environment
Table 4. Predicted environmental concentrations of manufactured nanoparticles in UK soil and waters
(from Boxall et al., 2007)
Particle type
Application
Water, µg/L
Soil, µg/kg
Aluminium oxide
Paint
0.002
0.01
Cerium dioxide
Scratch resistant coatings, catalysts
<0.0001
0.01
Fullerenes
Anti-inflammatory cream, eyeliner, face
powder, foundation, lipstick, mascara,
moisturizing cream, perfume
0.31
44.7
Gold
Face cream
0.14
20.4
Organosilica
Scratch resistant coatings
0.0005
0.07
Silver
Biocidal coatings, shampoo, soap,
toothpaste
0.01
1.45
Titanium dioxide
Paint, sunscreen
24.5
1030
Hydroxyapatite
Toothpaste
10.1
422
Latex
Laundry detergents
103
4310
Zinc oxide
Paint, scratch resistant coatings,
sunscreens
76
3190
Table 5. Comparison of UK exposure data for manufactured nanoparticles with toxicity data (from Boxall
et al., 2007)
Predicted in
water, µg/L
Toxicity data, µg/L
Other endpoints
Invertebrate
EC50
Fish
LC50
Algae
EC50
n-C60
0.31
>35,000
>>5000
-
Effects on invertebrate growth at
260 µg/L; bacterial growth
affected at 40µg/L; bacterial
phospholipids affected at 10 µg/L
TiO2
24.5
>100,000
>100,000
16,000
Effects on invertebrate growth at
2000 µg/L; bacterial growth
affected at 100,000 µg/L
SiO2
0.0007
-
-
No effect on bacterial growth at
500,000 µg/L
ZnO
76
-
-
No effect on bacterial growth at
100,000 µg/L
Fate of Manufactured Nanomaterials in the Australian Environment
31
Figure 10. Framework for deriving mass flow data for silver flows from nano-functionalised biocidal
plastics and textiles (from Blaser et al., 2008). Arrows represent silver flows; dashed lines indicate
different environmental spheres. TWT=thermal waste treatment; STP=sewage treatment plant.
The model predictions can be verified by comparison with measured data from different
aquatic environments.
Mueller and Nowack (2008) have followed a similar approach in the determination of
the expected exposure concentrations in air, soil and water for nanoparticulate silver and
titanium dioxide and for CNTs. Literature production data are used to determine the
weighted concentrations of nanomaterials from each product type. Release of
contaminants and their transfer between the various compartments in the model are then
determined using literature-derivations or best estimates of transfer coefficients. In
much the same way as that used by Boxall et al. (2007), PECs were derived for
particular environmental compartments, however, the number of product categories was
extended beyond the personal care and cosmetic products, to include all possible uses,
e.g. for silver, the categories were textiles, cosmetics, metal products, sprays and
cleaning agents, plastics, and paints. The findings are shown in Table 6.
The results were then compared with available toxicity data. No data were available for
soil toxicity. The EC50 value (concentration causing a 50% effect) used for silver
toxicity was 20-40 mg/L, but this was from bacterial toxicity testing (E. coli and
Bacillus subtilis), and so are not necessarily applicable. The authors noted that, for
ionic silver, literature LC50 values were 0.7 µg/L for algae and 2 µg/L for Daphnia.
They indicated that there was a lack of reliable toxicity data for TiO2. Their hazard
quotients (PEC/PNEC) indicate a potential concern for TiO2, compared to the
conclusions of Boxall et al. (2007) discussed above, but this may be a function of the
application of large assessment factors (1/1000) to the limited toxicity data.
Table 6. Predicted environmental concentrations (PEC) of nano-Ag, nano-TiO2 and CNTs in air, water
and soil (RE: realistic scenario; HE: high emission scenario) (from Mueller and Nowack, 2008)
32
Fate of Manufactured Nanomaterials in the Australian Environment
nano-Ag
nano-TiO2
CNT
Compartment
Unit
RE
HE
RE
HE
RE
HE
Water
µg/L
0.03
0.08
0.7
16
0.0005
0.0008
Water affected
by wastewater
µg/L
8
21
180
3933
na
na
Soil
µg/kg
0.02
0.1
0.4
4.8
0.01
0.02
Table 7. Hazard quotients (PEC/PNEC) for nano-Ag, nano-TiO2 and CNT in water (RE: realistic scenario;
HE: high emission scenario) (from Mueller and Nowack, 2008)
nano-Ag
Compartment
RE
HE
nano-TiO2
RE
HE
Water
0.0008
0.002
>0.7
>16
Water affected
by wastewater
0.2
0.5
>180
> 3900
a
CNT
RE
HE
0.005
0.008
naa
na
Not available
9.
9.1
ECOTOXICOLOGY OF NANOPARTICLES
Ecotoxicity and Nanoparticle Dose Metrics
In examining the ecotoxicity of nanoparticles to biological organisms, a critical question
is the determination of what influences the dose response. In traditional ecotoxicology
with soluble species, concentration is the dose measure, and specifically, the
bioavailable concentration, which may be some sub-set of the total concentration.
When dealing with nanoparticles, the concentration or mass metric may not apply, and
alternative considerations may involve particle number, surface area (shape), particle
composition, or surface reactivity.
In defining the applicable dose metric, some understanding of the mechanism of toxicity
of nanoparticles is required. Thus toxicity could be exerted by soluble species
dissociating from nanoparticles at a cell surface and crossing the cell membrane, or by
disruption of cell function by blockage of surface sites. If the nanoparticle is a
heterogeneous source of oxygen free radicals that are responsible for lipid peroxidation,
then it is likely that the dose will be dependent on the number of active sites on the
nanoparticles that are capable of free radical generation.
In studies of human toxicology of nanoparticles, there has been some debate about the
appropriate dose metric. Oberdorster et al. (2005) showed that surface area accounted
for differences in lung inflammatory effects of nanoparticulate TiO2 to rats and mice far
better than any mass considerations. Duffin et al. (2002) reached similar conclusions
for quartz nanoparticles, although noting the importance of surface reactivity.
Wittmaack (2007) disputed this interpretation, suggesting that particle number provided
Fate of Manufactured Nanomaterials in the Australian Environment
33
a better fit for differently prepared carbon nanoparticles, although the interpretation was
complicated by the possibility that aggregated particles might disaggregate on contact
with the lung. The relevance of these studies with atmospheric nanoparticles to toxicity
in aquatic or soil systems is, however, questionable.
Particle morphology may also be an important metric (Buzea et al., 2007). Particles can
be classified as having either high or low aspect ratios. The former include nanowires,
nanotubes and the like, while spherical, oval and cubic type particles have a low aspect
ratio. In pulmonary toxicology, particles with a high aspect ratio have been shown to be
more toxic (Inoue et al., 2006). The importance of aspect ratio in aquatic or terrestrial
toxicity is unknown.
No such definitive studies appear to have been undertaken for ecotoxicological
receptors, although there is clear evidence of greater toxicity of nanoparticulate versus
lesser surface area or macro forms, e.g. the demonstrated toxicity to water fleas
(Daphnia magna) of 30 nm TiO2 particles compared to no observed toxicity for 100500 nm aggregates of the same material (Lovern and Klaper, 2006).
To date, all toxicity data has been reported in terms of concentrations, but since
bioavailability will be dependent upon the physical properties, it will be necessary to
qualify all concentration data. Size is the next most critical parameter, since indications
are that when the nanoparticulate range is exceeded, properties approach those of the
parent macroparticles. Included in the size estimation should be a verification of the
fraction that is not in true solution, so that any observed effects are related only to the
particulate forms. It should be noted that the key environmental issue for any risk
assessment is to derive a no-effects concentration in the site-specific medium. Toxicity
determined in synthetic media might greatly over- (or under-) estimate toxicity because
of modification of bioavailability in the presence of colloids or other constituents.
The parallel in aquatic ecotoxicology to the consideration of site specific modifications
to water quality guidelines is a useful one (e.g. ANZECC/ARMCANZ, 2000). The
toxicity in synthetic media can be used to derive a conservative guideline trigger value
(for a given nanoparticle size) that might be modified by site specific chemistry. At this
stage of our knowledge, other physical and chemical measurements are probably
superfluous, given the already existing uncertainties in the measurements that are being
made.
A major practical issue with toxicity testing of nanomaterials is the dispersion of
nanoparticles in the test solutions. In aquatic toxicity testing, the contaminant is usually
in true solution and homogeneously distributed throughout the sample. Any attempts to
use artificial dispersants or sonication are likely to affect the degree of aggregation from
its natural state and so stirring of the sample is the only acceptable means of maintaining
the nanoparticles in any way close to a dispersed state, acknowledging that prolonged
stirring may also break up nanoparticles. Intermittent stirring might be an option.
A framework for nanoparticle toxicity assessment in waters based on the above
discussion is given in Table 8. A similar approach could be devised for testing in soils.
34
Fate of Manufactured Nanomaterials in the Australian Environment
Table 8. Approach to toxicity testing of nanomaterials in waters
Derivation of Nanomaterial Guideline Trigger Value
1.
Suspend nanomaterial in synthetic water (at an appropriate concentration with an
appropriate mixing time to achieve equilibrium solubility)
2.
Determine ‘soluble’ fraction (e.g. using dialysis)
3.
Determine insoluble fraction
4.
Determine particle size distribution on the sample from 1.
5.
Undertake toxicity tests using different species on the sample from 1 and on the
‘soluble’ fraction. Determine the contribution of ‘soluble’ species to the total
nanomaterial toxicity.
6.
Derive a guideline trigger value for the measured size distribution.
Derivation of a Site-specific Trigger Value
1.
Repeat the above approach using the appropriate site water.
Toxicity Testing of a Nanomaterial Sample for Comparison with Guideline Trigger
Values
1.
Repeat the above approach using either a synthetic or site water sample as
appropriate.
2.
Compare result with trigger value, noting compatibility of particle size
distribution.
Crane and Handy (2007) in a recent review of methods for characterising the
ecotoxicological hazard of nanomaterials suggested that rapid tests that identified
specific modes of toxicity, e.g. genotoxicity, immunotoxicity or oxidative stress assays
might be a useful addition to the standard suite of toxicity tests that uses algae,
invertebrates and fish. Because of the uncertainties in acute to chronic ratios in tests on
nanomaterials, it was recommended that where possible, chronic tests were preferable.
9.2
Toxicity to Aquatic Biota
9.2.1
Mechanisms of Biological Uptake and Toxicity
Studies in vitro at the cellular level point to oxidative stress as a key mechanism of
toxicity for many nanoparticles. Oxidative stress has been linked in a number of cases
Fate of Manufactured Nanomaterials in the Australian Environment
35
to the ability of many nanoparticles to generate reactive oxygen species (ROS: oxygen
ions, peroxides and free radicals) (Oberdorster et al., 2005, Nel et al., 2006).
Physical damage to cell membranes is also possible as a consequence of the abrasive
nature of some nanoparticles leading to toxicity (Stoimenov et al., 2002). Adhesion of
nanoparticles to the cell surface and dissociation of soluble toxic species can also
provide a route of uptake (Klaine et al., 2008; Apte et al., 2008).
Franklin et al. (2007) were unable to demonstrate algal cellular uptake of zinc from
nanoparticulate ZnO because of the unexpectedly high solubility of ZnO. They
subsequently demonstrated enhanced toxicity of CeO2 nanoparticles compared to bulk
CeO2 (Franklin et al., unpublished results), suggesting enhanced uptake.
For aquatic biota, nanoparticle uptake and potential toxicity will be dependent on the
type of organism, its trophic level and whether it is uni- or multicellular. With
unicellular organisms, the issue of whether nanoparticles can cross cell membranes
directly or via endocytosis is still a major question. For eukaryotic organisms, most
internalisation of nanoparticles will occur via endocytosis (Moore, 2006; Nowack and
Bucheli 2007), i.e. with the cell membrane enclosing the nanoparticles leading to their
deposition in the cytoplasm and association with intracellular organelles, without
directly passing through the cell membrane.
For higher organisms, uptake across the gill and other external surface epithelia is also
possible and interactions with aquatic plants may include adsorption onto the root
surface, incorporation into the cell wall, or diffusion into the intercellular space
(Nowack and Bucheli, 2007).
A further pathway for contaminant uptake is via the food chain. Direct ingestion is a
possibility for many organisms. Water fleas (Dapnia magna) rapidly ingested lipidcoated nanotubes via normal feeding behaviour, metabolizing the lipid coating as a food
source (Roberts et al., 2007). The toxic impact in many instances will depend on the
ability of the particles to promote cellular damage, e.g. by oxygen radical formation.
For example, SWCNTs observed in the gut lumen of fish exposed to sub-lethal
concentrations for 10-days, demonstrated an increase in oxidative stress markers and
ionoregulatory disturbance (Smith et al., 2007). More recently, direct evidence for a
dietary pathway of nanoparticle uptake has been demonstrated for uptake of quantum
dots in water fleas (Ceriodaphnia dubia) via a previously exposed algal food source
(Bouldin et al., 2008).
9.2.2
Ecotoxicity to Individual Species
Toxicity test data on manufactured nanomaterials from existing literature are
summarized in Table 9.
Bacterial toxicity
Many of the toxicity assessments of nanomaterials have focussed on bacteria, largely
undertaken using traditional growth media under optimum conditions. These data have
been well summarized elsewhere (Apte et al., 2008; Handy et al., 2008; Klaine et al.,
36
Fate of Manufactured Nanomaterials in the Australian Environment
2008). There is no doubt that many nanomaterials show bactericidal properties,
especially silver (Morones et al., 2005; Sondi and Salopek-Sondi, 2004), where that
property is the reason for its extensive usage. Similar antimicrobial activity is shown by
titanium dioxide (Wolfrum et al., 2002). More recent studies have demonstrated strong
antimicrobial activity of SWCNTs (Kang et al., 2007).
While these studies have been useful in investigating mechanisms of toxicity and
relative toxicities of different formulations (e.g. Lyon et al. 2005; Fang et al., 2007;
Yamamoto et al., 2001; Reddy et al., 2007), they will not be discussed in detail here: (i)
data from tests in growth media are not relevant to natural ecosystems; and (ii) bacterial
data are not used in species sensitivity distributions to determine safe concentrations of
nanomaterials in waters (DEW, 2007).
As already discussed in Section 5, the potential for impact on sewage bacteria is a
separate question to protecting organisms in natural waters.
Algal toxicity
Limited data are available for algal toxicity. The response to TiO2 is not particularly
sensitive (Hund-Rinke and Simon, 2006; Warheit et al., 2007) and that to ZnO is a
response to soluble zinc (Franklin et al., 2007).
Invertebrate toxicity
The freshwater crustacean Daphnia magna has been the most used invertebrate species
for nanomaterial toxicity testing. Daphnia were quite sensitive to n-C60 prepared by
tetrahydrofuran (THF) extraction (Zhu et al., 2006; Lovern and Klaper, 2006). It is
important to note that for these fullerenes, two preparation methods were followed, one
using sonication of fullerenes in water for 30 minutes to disperse the nanoparticles and
the second using the evaporation of THF from a THF extract added to water. The latter
were consistently more toxic to all organisms tested, and the question remains as to
whether the additional toxicity was due to THF, although these tests used THF only
controls. It has been suggested that sonication could enhance toxicity (Zhu et al., 2006).
Fish
Normally fish would be expected to show less sensitivity to dissolved contaminants than
algae or daphnids. This was not necessarily the case with nanomaterials, and may be
indicative of a different mechanism of toxicity, e.g. gill clogging, that would not occur
with dissolved contaminants.
Toxicity of nanoparticulate silver to zebrafish embryos has been demonstrated by Lee et
al., (2007). In this study, the only one to date of nanoparticulate silver toxicity to
aquatic biota, the endpoints were deformities and abnormalities in the embryos. No
EC50 values were quoted, but from the graphs were estimated to be in the range 10-20
ng/L. This is far lower than the bacterial toxicity values used by Mueller and Nowack
(2008) discussed earlier.
Fate of Manufactured Nanomaterials in the Australian Environment
37
The toxicity of soft nanoparticles has been poorly studied. The NICNAS (2000) report
on acrylic latex indicates that no toxicity data are available. They are generally believed
to have low toxicity.
38
Fate of Manufactured Nanomaterials in the Australian Environment
Table 9. Summary of toxicity testing results for manufactured nanomaterials (from Apte et al., 2008)
Nanomaterial
Size Fraction,
nm
Test Medium
Test species
Endpoint
Reference
n-C60 water-
Nominally 10-200
Standard USEPA medium
Water flea
Daphnia magna
48-h LC50 >35 mg/L
Zhu et al., 2006
Nominally 10-200
Moderately hard freshwater
USEPA protocol
Water flea
Daphnia magna
48-h LC50 0.8 mg/L
Zhu et al., 2006
n-C60 water-
Average diameter
Water flea
Daphnia magna
48-h LC50 7.9 mg/L
solubilised
30
Moderately hard freshwater
USEPA protocol
Lovern and Klaper
(2006)
n-C60 THF
10-20
Moderately hard freshwater
USEPA protocol
Water flea
Daphnia magna
48-h LC50 0.46 mg/L; NOEC
Lovern and Klaper
180 µg/L
(2006)
Synthetic hard water
Water flea
40% mortality at 2.5 mg/L over
Oberdorster et al., 2006
solubilised
n-C60 THFextract
extract
n-C60 water-
Nominal 10-200
Daphnia magna
solubilised,
21 days. No acute toxicity up
to 35 mg/L
n-C60 THF
Nominally 10-200
Standard USEPA medium
Fathead minnow
Pimephales promelas
0.5 mg/L 100% mortality in 618 h
Zhu et al., 2006
Nominally 10-200
Standard USEPA medium
Fathead minnow
Pimephales promelas
0.5 mg/L no effects after 48 h
Zhu et al., 2006
Nominally 30-100
Synthetic hard water
Juvenile largemouth bass
Mycropterus salmoides
0.8 mg/L 100% mortality in 618 h
Oberdorster, 2004
extract
n-C60 watersolubilised,
n-C60 THF
extract
Fate of Manufactured Nanomaterials in the Australian Environment
39
n-C60 water-
Nominally 10-200
Synthetic hard water
Freshwater
crustacea
Hyalella azteca
No toxicity below 7 mg/L
Oberdorster et al., 2006
Nominally 10-200
Synthetic hard water
Japanese medaka
Oryzias latipes
No acute toxicity at 0.5 mg/L
Oberdorster et al., 2006
solubilised,
n-C60 watersolubilised,
for 96h
Synthetic hard water
Zebrafish embryos
Danio rerio
1.5 mg/L was toxic
Zhu et al., 2007
?
Seawater
Meiobenthic
copepods
Amphiascus tenuiremis
Templeton et al., 2006
SWCNT as
prepared
?
Seawater
Meiobenthic
copepods
Amphiascus tenuiremis
No effects at 10 mg/L.
Evidence of ingestion and
aggregation
No effect at 1.6 mg/L; effects at
10 mg/L.
SWCNT
?
Freshwater with up to 0.15 mg/L
SDS
Rainbow trout
Oncorhynchus mykiss
Smith et al., 2007
SWCNT
?
Freshwater and seawater
Zebrafish embryos
Danio rerio
Effects on ventilation rate, gill
pathologies and gill mucus
secretion at 0.5 mg/L
Hatching delay at 150 mg/L
TiO2
Nominal 25
(small); 100
(large)
Average 140
Moderately hard water
OECD 201 protocol
Algae
Desmodesmus
subspicatus
Hund-Rinke and Simon,
2006
Moderately hard water
OECD 201 protocol
Algae
Pseudokirchneriella
subcapitata
Chlorophyll fluorescence
72-h EC50 44 mg/L small; no
dose response large
Chlorophyll fluorescence
72h EC50 16-21 mg/L
Nominal 25
(small); 100
(large)
30 THF; 100-500
sonicated
Moderately hard water
OECD 202 protocol
Water flea
Daphnia magna
No concentration-effect curve
observed up to 3 m/L
Hund-Rinke and Simon,
2006
Moderately hard freshwater
USEPA protocol
Water flea
Daphnia magna
48-h LC50 THF 5.5 mg/L;
Sonicated >500 mg/L
Lovern and Klaper,
2006
30 THF; 100-500
sonicated
Moderately hard freshwater
USEPA protocol
Water flea
Daphnia magna
No significant behavioural
changes LOEC 2.0 mg/L
Lovern et al., 2007
n-C60 THF
100 nm
extract
aggregates
SWCNT purified
TiO2
TiO2
TiO2 THF
dispersed;
sonicated
TiO2 THF
dispersed;
sonicated
40
Templeton et al., 2006
Cheng et al., 2007
Warheit et al., 2007
Fate of Manufactured Nanomaterials in the Australian Environment
TiO2
Average 140
Moderately hard water
OECD 202 protocol
Water flea
Daphnia magna
48-h EC50 >100 mg/L
Warheit et al., 2007
TiO2
24
De-chlorinated tap water
Rainbow trout
Oncorhynchus mykiss
Federici et al., 2007
TiO2
140
Moderately hard water
OECD 201 protocol
Rainbow trout
Oncorhynchus mykiss
No mortality during 14-day
exposure up to 1.0 mg/L.
Sub-lethal effects including gill
damage, observed.
96-h EC50 >100 mg/L
TiO2
TEM : 50 -400
De-chlorinated tap water
Carp
Cyprinus carpio
No mortality during 25 day
exposure to 10 mg/L.
Sun et al., 2007
TiO2
Nominal 19
De-chlorinated tap water
Carp
Cyprinus carpio
Zhang et al., 2007
ZnO
Average 178-361
USEPA, pH 7.5
Algae
Pseudokirchneriella
No mortality with 25 day
exposure to 10 mg/L TiO2.
Increased Cd accumulation
72-h EC50 68µg/L due to
dissolved Zn
8-day EC50 0.2-0.5 mg/L,
Adams et al., 2006
subcapitata
ZnO
Mixed
Spring water + food pellets
Water flea
Daphnia magna
Warheit et al., 2007
Franklin et al., 2007
possibly dissolved Zn
SiO2
Mixed
Spring water + food pellets
Water flea
Daphnia magna
8-day EC50 <10 mg/L
Adams et al., 2006
Cu
Nominally 80
De-chlorinated tap water
Zebrafish
Danio rerio
48-h LC50 1.5 mg/L
Griffit et al., 2007
Fe
Average 70
USEPA protocol
Water flea
Daphnia magna
48-h LC50 55 mg/L
Oberdorster et al., 2006
Ag
Average 12
Dilute NaCl
Zebrafish
Danio rerio
Embryo abnormalities EC50
Lee et al., 2007
10-20 ng/L
Quantum dots;
Estimated 10-25
Moderately hard water
Water flea
Ceriodaphnia dubia
96-h LC50 >110 µg/L
Bouldin et al., 2008
Cd/Se or Cd/Te
core with ZnS
Fate of Manufactured Nanomaterials in the Australian Environment
41
shell
Quantum dots;
Cd/Se or Cd/Te
core with ZnS
Estimated 10-25
Moderately hard water
Algae
Pseudokirchneriella
subcapitata
96-h LC50 37.1 µg/L of
Bouldin et al., 2008
quantum dots, estimated as 9.6
µg/L Cd and 2.4 µg/L Se
shell
42
Fate of Manufactured Nanomaterials in the Australian Environment
9.2.3
Developing Appropriate Guidelines for Nanomaterials in Waters
The available toxicity data are insufficient to develop reliable guidelines for most
nanomaterials in waters, however, it is instructive to attempt to derive low reliability
guidelines for the nanomaterials for which we have the most data. For the data from
Table 6 for n-C60 and TiO2, chronic NOEC values were obtained using a factor of 10 on
acute LC50 values or on EC50 values from acute endpoints (Table 10). Following
ANZECC/ARMCANZ (2000) guidelines, data would be required for an alga, an
invertebrate and a fish and the lowest NOEC would be then divided by a factor of 100.
In this case of n-C60, algal data are missing, however, if this is ignored, a value of 7.9
µg/L would be derived for the water-solubilised n-C60. A value for THF-extract n-C60 is
more problematic and clearly <0.5 µg/L. The OECD approach would use a factor of
1000 on the lowest NOEC. For TiO2 the lowest result is for a THF-extracted sample.
Ignoring that, the PNEC for TiO2 dispersed by sonication would be 40 µg/L.
Table 10. Data for estimation of guideline concentrations for n-C60 in freshwater
Nanomaterial
Formulation
Species
Endpoint, mg/L
n-C60
Water
solubilised
by sonication
Daphnia magna
48-h LC50 7.9
0.79
n-C60
Water
solubilised
by sonication
Pimephales
promelas
No effects after
48 h at 0.5
>0.05
n-C60
THF extract
Daphnia magna
48-h LC50 0.8,
0.46 (acute
NOEC 180 µg/L)
0.08, 0.05
n-C60
THF extract
Pimephales
promelas
100 % mortality
in 6-18 h 0.5
<0.05
n-C60
THF extract
Mycropterus
salmoides
100% mortality
in 6-18 h 0.8
<0.08
n-C60
THF extract
Danio rerio
<1.5
<0.15
TiO2
No THF
Desmodesmus
subspicatus
72-h EC50 44
8.1
TiO2
No THF
Pseudokirchneriella
subcapitata
72-h EC50 1621
4.0
TiO2
THF extract
Daphnia magna
48-h LC50 THF
5.5
0.55
TiO2
No THF
Daphnia magna
>500
>50
Fate of Manufactured Nanomaterials in the Australian Environment
Estimated
chronic NOEC,
mg/L
43
With this extra conservatism, the PNEC value for both n-C60 and TiO2 are seen to be
now getting closer to the PEC values estimated in the UK (Table 5), with all of their
limitations. This highlights, if nothing else, the need for additional toxicity data.
9.2.4
Bioaccumulation
The published evidence to date for the bioaccumulation of manufactured nanomaterials
by aquatic organisms is limited and is summarised in Table 11. There is TEM evidence
of the presence, in the cytoplasm of bacterial cells (e.g. Escherichia coli, Bacillus
subtillus, Staphylococcus aureus), of MgO (Makhulf et al., 2005), SWCNTs (Kang et
al., 2007), ZnO (Brayner et al., 2006), quantum dots (Kloepfer et al., 2005) and silver
(Xu et al., 2004; Morenes et al., 2005). Many of these studies indicated cellular
damage, however, intracellular uptake was only indicated for MgO (<11 nm) and silver
nanoparticles (<80 nm), and for quantum dots (<5 nm).
As noted earlier, an important finding was the food chain transfer of quantum dots via
exposed algae to water fleas (Bouldin et al., 2008). The quantum dots have a CdSe core
and a ZnS shell. The coatings appeared to provide protection from toxicity to cadmium
(or selenium), but transfer of core metals from intact nanocrystals occurred at levels
well above toxic thresholds to the water fleas.
Table 11. Published evidence of nanoparticle uptake by aquatic organisms (expanded from Apte et al.,
2008)
Nanoparticle
Organism
Target
Organ
Evidence
Reference
Bacteria
MgO
Escherichia coli
Bacillus
subtillus
Membrane
TEM images confirm damage
and leakage of cell contents.
Stoimenov,
2002
SWCNT
Escherichia coli
Membrane
Increased membrane
permeability in cells in direct
contact with SWCNT.
Physical damage to the
membrane and leakage of cell
contents is proposed.
Kang et al.,
2007
MgO
Escherichia coli
Staphylococcus
aureus
Whole cell
TEM shows ultrastructural
changes on exposure to 8±1
and 11±1 nm particles.
Elevated intracellular Mg
confirmed.
Makhluf et al.,
2005
ZnO
Escherichia coli
Whole cell
TEM reveal electron dense
areas in the cytoplasm. No
elemental analysis.
Brayner et al.,
2006
Quantum dots
Escherichia coli
Whole cell
TEM, fluorescence
spectroscopy show adenine-
Kloepfer et al.,
44
Fate of Manufactured Nanomaterials in the Australian Environment
Bacillus subtilis
conjugated QDs < 5 nm are
internalised. Intracellular Cd
and Se confirmed.
2005
Ag
Pseudomonas
aeruginosa
Whole cell
Particles up to 80 nm
transporting in and out of
cells. TEM images confirm
electron dense areas in the
cytoplasm.
Xu et al., 2004
Ag
Escherichia coli
Whole cell
TEM images showing
electron dense intracellular
areas. EDS elemental
mapping confirms Ag
distribution throughout the
cell. 1 -10 nm particles
interact preferentially with
the cell.
Morones et al.,
2005
SWCNT lipid- Daphnia magna
coated
Gut
Rapid (45-min) ingestion and
presence of lipid-coated
SWCNT in the gut track
observed in time-course
micrographs.
Roberts et al.,
2007
Quantum dots
Ceriodaphnia
dubia
Gut
Evidence for food chain
transfer of core metals from
quantum dot-dosed algae
Bouldin et al.,
2008
Cu
Danio rerio
Gill
Histopathological analysis
revealed damage to gill
lamellae by proliferation of
epithelial cells and oedema of
gill filaments. Unclear if
effects mediated by particle
uptake.
Griffitt et al.,
2007
TiO2
Oncorhynchus
mykiss
Gill
Histopathological changes to
the gill and gut but fish did
not accumulate TiO2 in the
internal organs.
Federici et al.,
2007
Oncorhynchus
mykiss
Gill
Histopathological changes to
the gill and gut and liver.
Aggregated SWCNTs
observed in the gut lumen.
Smith et al.,
2007
Fish
SWCNT
9.2.5
Gut
Gut
Ecological Impacts
There have been no published studies on the broader ecological impacts of
manufactured nanoparticles.
Fate of Manufactured Nanomaterials in the Australian Environment
45
9.3
Sediment Toxicity
Given that sediments are the ultimate receptor of nanoparticles in aquatic systems,
benthic organisms are likely to be as big a concern as those in the overlying water. The
nanoparticles are likely to be highly aggregated in the sediments, so any unique toxic
properties associated with nano size are likely to be absent. Very few studies have
looked at nanomaterials in sediments. For example, Kennedy et al. (2008) showed that
the survival of several amphipods was affected by MWCNTs in whole sediment
bioassays, but at unrealistic concentrations exceeding 100 g/kg. More studies are
required to fully assess nanoparticle properties (aggregation, surface area),
bioavailability and toxicity in the more complex sediment environment.
9.4
Toxicity to Terrestrial Biota
9.4.1
Ecotoxicity to Individual Species
There are very few data by which to assess the potential environmental risk of
nanoparticles to the terrestrial environment and this is seen as a key knowledge gap by
regulators (US EPA , 2007). As yet, there are few reports in the peer-reviewed
scientific literature of the assessment of ecotoxicity of nanoparticles to soil biota, in
soils. Several reports have examined ecotoxicity to soil organisms, but the media used
have been simple aqueous media (Brayner et al., 2006; Yang and Watts, 2005; Zheng et
al., 2005; Lin and Xing, 2007) and persistence of the nanoparticles in the test media was
not assessed. These are summarized in Table 12.
Table 12. Toxic effects of nanomaterials on soil organisms (from Klaine et al., 2008).
Nanomaterial
Toxic Effects
References
None. Endpoints tested were respiration (basal and
substrate-induced), microbial biomass C, enzyme
activities. Small shift in bacterial and protozoan
gene patterns by PCR-DGGE.
Tong et al., 2007
No effect on respiration (basal), microbial biomass
C (measured by substrate-induced respiration) and
protozoan abundance. Reduction in numbers of
bacteria. Small shift in bacterial and protozoan
gene patterns by PCR-DGGE.
Johansen et al.,
2008
No effect on seed germination and root growth of
corn, cucumber, lettuce, radish, and rape. Reduced
Lin and Xing,
2007
Carbon-containing
A) Fullerenes
C60 granular and
C60 water suspension (nC60)
n-C60
B) Carbon nanotubes
Multi-walled
46
Fate of Manufactured Nanomaterials in the Australian Environment
root growth of ryegrass.
Metals
Aluminium
No effect on seed germination of corn, cucumber,
lettuce, radish, rape, and ryegrass. Rhizotoxic to
corn, lettuce, and ryegrass but stimulated radish and
rape root growth.
Lin and Xing,
2007
Zinc
Reduced seed germination of ryegrass and reduced
root growth of corn, cucumber, lettuce, radish,
rape, and ryegrass
Lin and Xing
2007
Phytotoxic (germination and seedling growth) but
see text.
Yang and Watts,
2005
No effect on seed germination of corn, cucumber,
lettuce, radish, rape, and ryegrass. No effect on root
growth of cucumber, lettuce, radish, rape, and
ryegrass. Reduced root growth of corn.
Lin and Xing,
2007
TiO2
Stimulatory to spinach seed germination and
seedling growth at low dose, phytotoxic at high
doses
Zheng et al., 2005
ZnO
Reduced seed germination of corn and reduced root
growth of corn, cucumber, lettuce, radish, rape, and
ryegrass
Lin and Xing,
2007
Metal oxides
Al2O3
Yang and Watts (2005) reported the toxicity of alumina nanoparticles (13 nm, coated
with and without phenanthrene) to root growth of five plant species (cabbage, carrot,
corn, cucumber, and soybean) exposed to aqueous suspensions of the nanoparticles, but
only at high concentrations (2,000 mg/L). Loading of the alumina nanoparticles with
phenanthrene reduced the toxicity of the nanoparticles. The nanoparticles were not
physically characterised prior to dosing, doses were not analytically confirmed, and in a
letter to the Editor of Toxicology Letters, Murashov (2006) pointed out the experimental
protocol of Yang and Watts (2005) did not distinguish toxicity caused by application of
the aluminium in a nanoparticle form, and toxicity of solution aluminium derived from
the nanoparticle. Indeed aluminium is a major component of soil minerals, known to be
phytotoxic in acidic soils for almost a century (Magistad 1925) so the phytotoxicity
observed by Yang and Watts (2005) is not surprising, and clearly indicates the need to
accurately determine if the nanoparticulate form of a contaminant is toxic, or if the
soluble contaminant derived from the nanoparticle is toxic. Franklin et al. (2007)
reached similar conclusions for the toxicity of ZnO nanoparticles to aquatic biota.
Zheng et al. (2005) examined the effects of nano- and bulk-TiO2 on spinach seed
germination and early plant growth in simple Perlite media containing a complete
nutrient solution. Nano-TiO2 significantly increased seed germination and plant growth
at low concentrations, but decreased these parameters at high concentrations. Bulk-
Fate of Manufactured Nanomaterials in the Australian Environment
47
TiO2 had little effect. The manufactured nanoparticles in this study were not physically
characterised and no details of size or surface reactivity of the materials were provided.
Recently, Lin and Zing (2007) examined the toxicity of several nanoparticles
(MWCNTs, Al, Al2O3, Zn and ZnO) to germination and early root growth of six plant
species in simple aqueous media at pH 6.5–7.5. The nanoparticles were not physically
characterised prior to exposure and doses were not confirmed. The zinc-based
nanoparticles had the greatest effect on plant germination and root growth, with EC50
concentrations similar for both zinc- and ZnO-nanoparticles of 20–50 mg/L depending
on plant species. The authors attempted to quantify the solution zinc dose in their
experiments by centrifugation (3000 G for 60 min) and filtration (0.7 μm). They
reported that the centrifugation procedure did not fully separate the nanoparticles from
the solution phase (assessed using TM-AFM), but they did not provide microscopic
information on the solutions after filtration. Surprisingly, a 2000 mg/L suspension of
ZnO after centrifugation and filtration returned a solution zinc concentration of only
0.3-3.6 mg/L, significantly less than the concentration of Zn2+ in equilibrium with bulk
ZnO at pH 6.5–7.5, ~10–900 mg/L (Lindsay, 1979). Copper nanoparticles were also
recently found to be potentially phytotoxic (Lee et al., 2008).
To date, there are only two reports in the literature of the terrestrial effects of
nanoparticles performed in soil, both on fullerenes (Tong et al., 2007; Johansen et al.,
2008). Tong et al. (2007) examined the toxicity of n-C60 in aqueous suspension and in
granular form to soil microorganisms using soil respiration, microbial biomass,
phospholipid fatty acid analysis, and enzyme activities as endpoints. The authors also
examined the DNA profile of the microbial community. All tests were performed in the
laboratory at optimal moisture conditions. In contrast to the observed microbial toxicity
of n-C60 in vitro (Fortner et al., 2005), Tong et al. found no effect of n-C60 to any
endpoint in the soil medium used (silty clay loam, 4% organic matter, pH 6.9). They
suggested that this was due to the strong binding of n-C60 to soil organic matter,
although no evidence was provided that organic matter was the solid phase in soil
reducing the effective dose. A similar set of experiments was performed by Johansen et
al. (2008), who examined the effect of n-C60 added to a neutral soil (pH 6.7) with low
organic C content (1.5%) on soil respiration, biomass C, bacterial and protozoan
abundance and the PCR-DGGE profiling of bacterial and protozoan DNA. No effects of
exposure of n-C60 were found on soil respiration, biomass C, and protozoan abundance,
but reductions in bacterial abundance were observed through colony counts. The n-C60
also caused only a small shift in bacterial and protozoan DNA, indicating a small
change in community structure, similar to the results of Tong et al. (2007). Similar
results from the same group were recently published for anaerobic bacteria typical of
wastewater sludge treatment systems (Nyberg et al. 2008).
There have been few reports of bioaccumulation or trophic transfer of nanomaterials to
soil invertebrates or mammals. A recent study of bioaccumulation of SWCNTs by
earthworms indicated a very low bioaccumulation factor compared to pyrene (~100-fold
lower) (Petersen et al., 2008), and a study of TiO2 accumulation by isopods (Porcellio
scaber) also indicated a low bioaccumulation potential for these nanomaterials (Jemec
et al., 2008).
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These data highlight the need for more information on the interaction of nanoparticles
with soil components, and more quantitative assessments of aggregation/dispersion,
adsorption/desorption, precipitation/dissolution, decomposition and mobility of
manufactured nanoparticles in the soil environment. This information will aid the
interpretation of terrestrial ecotoxicity test data, and will inform the correct protocols for
the assessment of the ecotoxicity of nanoparticles in soils.
9.4.2
Development of Guidelines for Nanomaterials in Soils
The available toxicity data are insufficient to develop reliable guidelines for most
nanomaterials in soils. Effects have been inconsistent and studies of high quality have,
to date, not demonstrated significant adverse effects when soil was the medium used for
testing. It is therefore premature to suggest any regulatory limit for any nanomaterial in
soils.
10. INTERNATIONAL PROGRESS ON NANOPARTICLE RISK
ASSESSMENT
10.1
International Approaches
With nanotechnology industries growing exponentially worldwide, the assessment of
the risks they pose to the environment is still being pursued by government agencies.
Although it is recognised that available toxicity data on macro-sized chemicals will not
necessarily apply at the nanoscale, the current approach is still largely one of
information gathering through funding of additional research and development that will
provide a more sound basis than currently exists for managing the environmental
impacts of manufactured nanomaterials.
The field is evolving extremely rapidly, and it is important to regularly check the
literature. CSIRO are part of an international Nanoparticles Advisory Group in the
Society of Environmental Toxicology and Chemistry that shares on a monthly basis the
latest research and regulatory developments, while the Nanosafety Theme in CSIRO’s
Niche Manufacturing Flagship has close links with Dr Andrew Maynard of the
Woodrow Wilson International Centre for Scholars (see below). Such linkages are vital
to both contributing to and accessing the latest information. CSIRO also has links into
the OECD Working Party on the Safety of Manufactured Nanomaterials, as discussed
later.
10.1.1 USA
In the US, a National Nanotechnology Initiative (NNI, 2001) was launched by the
National Science and Technology Council in 2001. Funding was provided to support
nanoscience and technology research via a range of major agencies (e.g. NSF, NIH,
DOE, NASA, NIST, EPA, etc.,) in a number of different theme areas. Environmental
issues were only of marginal concern. The National Science Foundation (NSF) later
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49
established six facilities as part of Nanoscale Science and Engineering Centers. The
Center for Biological and Environmental Nanotechnology at Rice University was the
facility focussing on environmental issues (CBEN, 2005).
The Woodrow Wilson International Centre for Scholars and the Pew Charitable Trusts,
based in Washington DC, established a Project on Emerging Nanotechnologies in 2005.
This project has had a leading input to the nanotechnology debate in the US and beyond
(PEN, 2007b). Its publications (e.g. Maynard, 2006; Greenwood, 2007) have been a
vehicle for some useful basic information. Its inventory on nanoparticle usage (PEN,
2007a) is particularly valuable.
The US Environmental Protection Agency (US EPA) has been coming to grips with
how to apply the Toxic Substances Control Act to nanotechnology (Greenwood, 2007).
A report prepared by the Woodrow Wilson Institute for Scholars investigated the
dilemmas facing manufacturers and the USEPA in trying to deal with nanomaterials
under that Act, using as an example, carbon nanotubes (WWIS, 2003). There were
many uncertainties as to whether management in this way would be effective.
Nevertheless, the USEPA recently successfully fined a technology company over
$200,000 for selling unregistered nanopesticides (PEN, 2007b). The fine was made
under the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA).
A nano risk framework was prepared in 2007 in a partnership between the
Environmental Defense Fund and DuPont (Environmental Defense-DuPont, 2007). The
framework identified a basic set of environmental fate data including nanomaterial
aggregation and disaggregation in the exposure media and screens for persistence and
biodegradability. For exposure assessments, they recommended acute toxicity and
bioaccumulation testing, but identified a need for ecosystem level studies of effects on
populations. Chronic tests would be required if a nanoparticle was potentially persistent
and bioaccumulative. Depending on the fate, sediment testing might also be triggered.
The USEPA published a definitive Nanotechnology White Paper in 2007, following a
three-year review, to inform EPA management of the science needs associated with
nanotechnology. It included recommendations for addressing science issues and
research needs (USEPA, 2007). More recently they produced a Draft Nanomaterial
Research Strategy to guide the nanotechnology research program within the EPA’s
Office of Research and Development (USEPA, 2008). They identified four key research
themes and seven key scientific questions which highlight the limitations of our current
knowledge:
1. Sources, Fate, Transport and Exposure
a. Which nanomaterials have a high potential for release from a life-cycle
perspective?
b. What technologies exist, can be modified, or must be developed to detect
and quantify engineered nanomaterials in environmental media and
biological samples?
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c. What are the major processes/properties that govern the environmental
fate of engineered nanomaterials, and how are these related to physical
and chemical properties of these materials?
d. What are the exposures that will result from the releases of engineered
nanomaterials?
2. Human Health and Ecological Research to Inform Risk Assessment and Test
Methods
a. What are the effects of engineered nanomaterials and their applications
on human and ecological receptors, and how can these effects be best
quantified and predicted?
3. Risk Assessment Methods and Case Studies
a. Do Agency risk assessment approaches need to be amended to
incorporate special characteristics of engineered nanomaterials?
4. Preventing and Mitigating Risks
a. What technologies or practices can be applied to minimize risks of
engineered nanomaterials through their life cycle, and how can
naonotechnology’s beneficial uses be maximised to protect the
environment?
10.1.2 United Kingdom
The Royal Society and Royal Academy of Engineering released a report in 2004 on
nanoscience and nanotechnologies that addressed the current state of environmental
assessment of nanomaterials. It proposed that nanoparticulate forms of chemicals
should be treated as new chemicals for regulatory purposes, and identified the need for
new research to determine routes of exposure and toxicity. The UK government has
released several reports investigating the potential risks posed by manufactured
nanoparticles (DEFRA, 2005, 2007). The reports place the UK research program
overseen by a cross-government Nanotechnology Research Coordination Group in an
international context. They are collaborating with the OECD and the International
Standards Organisation (ISO) to share data and experiences to maximise the speed with
which potential risks can be identified and managed.
Specific task forces are addressing: (i) Metrology, characterisation, standardisation and
reference materials, (ii) Exposures: sources, pathways and technologies, (iii) Human
health and hazard assessment, (iv) Environmental hazard and risk assessment, and (v)
Social and economic dimensions of nanotechnologies.
A regulatory gaps analysis undertaken by Frater et al. (2006) for the UK Department of
Trade and Industry identified a number of gaps in the application of environmental
regulations to nanomaterials. A lack of knowledge of toxicity data was a critical issue.
Fate of Manufactured Nanomaterials in the Australian Environment
51
10.1.3 Other International Activities
The OECD’s Environment Directorate has been active in sponsoring a number of
meetings dealing with the safety of manufactured nanomaterials. Details of these are
available on their website (http://www.oecd.org/ehs). Reports on developments in
China, Japan, Italy and Germany were included at the 2005 workshop in Washington.
The OECD established a Working Party on the Safety of Manufactured Nanomaterials
(WPMN) in 2006.
Eight steering groups (SG) have been established within the WPMN to run the
following projects:

SG1: Development of an OECD database on EHS research

SG2: EHS Research Strategies on Manufactured Nanomaterials

SG3: Safety Testing of a Representative Set of Manufactured Nanomaterials

SG4: Manufactured Nanomaterials and Test Guidelines

SG5: Co-operation on Voluntary Schemes and Regulatory Programmes

SG6: Co-operation on Risk Assessments

SG7: The Role of Alternative Methods in Nanotoxicology

SG8: Exposure Measurement and Exposure Mitigation.
At the OECD WPMN Workshop in Tokyo in April 2007, attended by Drs Maxine
McCall and Simon Apte of CSIRO and NICNAS staff, a sponsorship program was
initiated whereby member countries volunteered to undertake work on specific
nanomaterials of national interest in collaboration with each other. Australia agreed to
undertake the study of zinc oxide, cerium dioxide and silver.
Nanoparticles are fully covered by REACH, the new European Community regulation
on chemicals and their safe use requirements. One of the first activities of the member
State Committee under the European Chemicals Agency (ECHA) was to institute a
Nanoparticles Working Group. Nominations for this Working Group were sent to
ECHA by member states and observers from industry and other countries including the
US. .
The European Chemical Industries Council (CEFIC) is currently reviewing strength and
weaknesses of the REACH risk assessment framework for nanoparticles, building on
the EU SCENIHR (Scientific Committee on Emerging and Newly Identified Health
Risks) report which covers the nanoparticles risk assessment topic (SCENIHR, 2005).
A summary of activities in Canada as of 2005, (Bergeron and Archambault, 2005)
indicates a similar scarcity of information, and identified needs for research and data
collection and the need to benefit from European and US experiences.
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10.2
Australian Activities
Australia has been following a path similar to other major international players in its
approach to nanomaterial risk assessment. NICNAS, the national regulator of industrial
chemicals, issued a voluntary call for information to importers and manufacturers of
nanomaterials in 2006, as discussed earlier. Industry was asked to provide information
on uses and quantities of nanomaterials imported or manufactured for industrial
purposes, including use in cosmetics and personal care products. The information was
designed to assist in understanding which nanomaterials are available on the market or
close to commercialisation, and help focus our efforts to ensure the adequacy of the
regulatory scheme to assess nanomaterials.
Nanomaterials used exclusively as therapeutic goods, pesticides or food additives do not
fall within the scope of NICNAS, and were consequently outside the voluntary call for
information. The results of its findings were published on its website
http://www.nicnas.gov.au. Information from a new call is currently being compiled.
Chemicals that are not listed on the Australian Inventory of Chemical Substances
(AICS), which is based on the chemical formula and CAS number of chemicals (with no
size definition), are generally regarded as "new" and must be notified to NICNAS and
assessed for human health and environmental risks prior to their introduction and use.
Nanoscale forms of chemicals already listed on AICS (i.e. having an identical chemical
formula and CAS number) are currently considered to be "existing" chemicals. These
nanoscale existing chemicals can be selected for assessment if they potentially present a
changed risk of adverse health and/or environment effects. To date, NICNAS has not
assessed any nanomaterials with novel properties.
NICNAS is currently examining the suitability of its regulatory framework and
processes to protect human health and the environment in association with the OECD
WPMN, and by engagement with Australian government agencies under the National
Nanotechnology Strategy. At the same time, NICNAS has convened a Nanotechnology
Advisory Group which has three members each from the community and industry, and
two members from academia and one from NICNAS, and which NICNAS chairs.
A national Nanotechnology Roundtable was hosted by the National Health and Medical
Research Council (NHMRC) in December 2006. The Roundtable was attended by
representatives from Australian academic institutions, health and environment
government departments, regulatory bodies and industry and a representative from the
New Zealand Health Research Council. The Chief Executive Officer of the NHMRC is
using the outcomes of the Roundtable to inform future directions for the NHMRC.
In June 2006, the National Nanotechnology Strategy Taskforce produced a report for the
government on "Options for a National Nanotechnology Strategy" (NNST, 2006). Its
major findings with respect to the environment were:

nanomaterials do exhibit novel properties that will have health safety and
environmental implications, but the significance is unclear at the moment, and
cannot be easily predicted due to a gap in knowledge;
Fate of Manufactured Nanomaterials in the Australian Environment
53

measurement and assessment of nanoparticles is a key priority for further
research;

there is a need for continued international cooperation in the field;

the Australian regulatory system needs to be flexible to address the challenges
posed by nanoparticles, and that coordinated processes at the regulatory level
need to be put in place; and

while no serious risk is evident now, potential real risks in the use of
nanotechnologies in Australia must be identified so that appropriate risk
management strategies can be employed for their safe use.
Following on from the Taskforce’s report, a HSE Working Group comprising federal
agencies with responsibility for policy and implementation of Australia's regulatory
frameworks was established to consider HSE issues in more detail. This group
commissioned a review of the capacity of Australia's regulatory frameworks to manage
any potential impacts of nanotechnology, which was produced in 2007 (Ludlow et al.,
2007).
Also in 2006, the TGA conducted a review of the scientific literature in relation to the
use of nanoparticulate zinc oxide and titanium dioxide in sunscreens, concluding that
they did not represent a major health threat. At that time, Food Standards Australia
New Zealand (FSANZ) had not received any applications to consider the regulation of
any nanomaterials under the Australia New Zealand Food Standards Code.
The APVMA have also recently published in the Gazette, a voluntary call for
information on nanomaterials in agricultural or veterinary chemicals, or agricultural and
veterinary chemical products. APVMA has published a position paper on
nanotechnology (APVMA, 2008).
The Department of Infrastructure, Transport, Regional and Local Government (transport
of hazardous materials) and the Department of the Environment, Water, Heritage and
the Arts (DEWHA) in conjunction with the above bodies are currently assessing
international research in this area.
11. DEVELOPMENT OF TECHNICAL GUIDELINES FOR
NANOMATERIAL ASSESSMENT
An Environmental Risk Assessment Guidance Manual for industrial chemicals was
issued by the then Department of Environment and Water Resources (now DEWHA) in
2007 (DEW, 2007). While this did not consider nanomaterials, the draft framework
outlined an approach that was consistent with NChEM, the discussion paper on a
national framework for chemicals management in Australia prepared by the
Environment Protection and Heritage Council (EPHC, 2006).
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11.1
Exposure Assessment Incorporating Nanomaterial Fate
The development of appropriate technical guidelines that cover the impacts of
manufactured nanomaterials in waters, sediments and soils first requires an evaluation
of the fate of nanomaterials in these environments. Only then can exposure parameters
be appropriately assessed for comparison against any available environmental quality
guidelines. It is evident from the information presented in this report, that the fate of
nanomaterials in the environment, and their toxicity to biota, is likely to be a function of
size, shape, surface properties and bulk composition. Inferring fate and toxicity from
bulk composition alone (e.g. zinc oxide, CNTs, nanodots) is likely to be inappropriate
due to the wide range of surface functional coatings applied to many nanomaterials.
Surface properties will therefore play a key role in interactions with all environmental
matrices (soils, sediments and waters).
It is possible to lay out the following key questions that incorporate basic considerations
of nanomaterial fate in the form of a required check list:
A. Nanomaterial Classification
(i)
What is the class of the nanomaterial (e.g. metal oxides; carbon products (nC60 fullerenes, CNTs; metals; quantum dots and semiconductors; nanoclays;
dendrimers, and nanoemulsions)?
(ii)
What is its core chemical component (e.g. zinc oxide, silver, SWCNT etc)?
(iii)
Is the basic formulation modified by additives?
(iv)
What is the nominal particle size of the solid phase component?
B. Fate in Waters
The following considerations are required if the nanomaterials are to enter aquatic
systems, noting that the behaviour may differ for different product formulations:
(i)
What is the particle size of dispersed nanomaterials in a natural receiving
water system (i.e. extent of aggregation) as determined by appropriate
techniques? The key factor here is the existence of dispersed or aggregated
particles that are in the nano range (<100 nm) and likely to have properties
differing from equivalent bulk macroparticles.
(ii)
Does this particle size change with time, and if so over what timescale
(hours, days, weeks)?
(iii)
What fraction of the nanomaterial dispersion is soluble (as determined by
dialysis or ultrafiltration) and/or dissociated thereby having potentially
different biological availability to the insoluble fraction?
(iv)
Does the soluble fraction change with time, and if so over what timescale?
Fate of Manufactured Nanomaterials in the Australian Environment
55
(v)
What is the estimated mass concentration of the insoluble aggregated
dispersion in the nano size range? If the concentration is low, are
interactions with natural colloids at higher concentrations likely to modify
the form of the nanomaterial through adsorption.
C. Fate in Sediments
If nanomaterials accumulate in sediments, it would be assumed that this is the
consequence of settling from the water column due to a loss of buoyancy. To do so
requires considerable aggregation of particles or association with other water-borne
particulates notionally to exceed 1 µm in size, although this is density dependent. Once
accumulated in the sediments, the key fate questions relate to dissolution and
transformation. The key questions might be:
(i)
Is there any dissolution of nanomaterials in the sediment pore waters under
the redox, pH and microbial conditions existing in the sediments?
(ii)
Does this soluble fraction change with time?
(iii)
What is the particle size of the nanomaterials in the sediment, i.e. are there
any nano-sized manufactured particles that might pose a different threat to
natural nanoparticles, or natural or manufactured macroparticles.
D. Fate in Soils
As already noted, nanomaterial interactions in soils are poorly characterised. In terms of
key questions, similar issues to those discussed above will need to be considered:
(i)
Is the nanomaterial soluble in soil pore waters and so able to exert effects in
that form? How is this solubility affected by soil pH, salinity, sodicity, redox
conditions and time?
(ii)
What is the particle size of nanomaterials in soil after any natural
aggregation processes? How easily are nanomaterials sorbed and retained by
soil minerals and organic matter?
(iii)
Are nanomaterials more mobile through soils than natural nanoparticles?
Can they be vectors for enhanced transport of contaminant solutes to
groundwaters, e.g. pesticides, metals, dioxins, etc?
11.2
Effects Assessment Incorporating Nanomaterial Fate
The information obtained from the above check list provides the necessary input to the
second component of any hazard assessment, the effects assessment. It will dictate the
evaluation of potential toxicity based on the existing toxicity database and a knowledge
of how changes to basic nanomaterials in the environment affect their bioavailability.
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In terms of defining guideline concentrations for nanomaterials in waters, sediments and
soils, data are required from an appropriate set of toxicity tests (DEW, 2007) that can be
used to derive an appropriate PNEC. For aquatic biota, data are needed from at least
five species from four taxonomic groups if a statistical extrapolation method is to be
applied (ANZECC/ARMCANZ, 2000), although this size dataset must be viewed as a
minimum. Alternatively a low reliability value can be estimated by the application of an
assessment factor to the lowest chronic no observed effects concentration (NOEC). For
water sample toxicity data, as shown in Table 10, this will currently be the only
alternative.
The limitations of these toxicity datasets have been discussed earlier, the principal
concern being that the nanomaterials used are appropriately characterised in terms of
key parameters, especially particle size (degree of aggregation), and any specific
formulations that may modify behaviour (e.g. surfactant additions or surface coatings).
In addition, there is a need for site-specific, or at best water-, sediment- or soil-specific
guidelines that take into account environmental chemistry and its effect on nanoparticle
behaviour (especially the effect of ionic strength on aggregation).
Based on findings to date, the following hypotheses are suggested with respect to the
bioavailability and potential toxicity of nanomaterials in the environment:
(i)
The fate and bioavailability of a particular nanomaterial might be expected to
change in the presence of additives that affect surface properties;
(ii)
Small dispersed nanoparticles (<20 nm) are likely to be more bioavailable
and potentially toxic than large aggregates (>100 nm). Aggregates can
nevertheless exert toxicity compared with bulk material especially in
instances related to ROS generation where the aggregate surface area may be
only marginally lower than that of its composite nanoparticles.
(iii)
Interaction with other particles and aggregation will be dependent on both
the surface charge and the surface area of nanoparticles. Low surface area
and net negatively charged particles are less prone to aggregation and are
potentially more mobile (but perhaps less bioavailable) since most
membranes have negative trans-membrane potentials;
(iv)
Interaction of nanoparticles with other naturally present colloids or organic
macromolecules will also affect reactivity. Such interactions will be
favoured by high nanoparticle surface areas and excess concentrations of
natural colloids, and would be expected to result in larger particles with
reduced bioavailability;
(v)
Aggregation is faster in environments of high ionic strength, i.e. high
hardness or saline waters, saline soils, or saline sediments;
(vi)
With greater aggregation, particle toxicity approaches that of the equivalent
bulk macroparticles;
Fate of Manufactured Nanomaterials in the Australian Environment
57
(vii)
For metallic nanomaterials, water-soluble and potentially dissociated
nanoparticles are likely to be more toxic than insoluble and undissociated
particles of the same material, although there are mechanisms by which
insoluble materials can exert toxicity; and
(viii) The kinetics of both aggregation and dissolution will influence the above
toxicity.
Current indications are that it will only be possible to provide low reliability PNECs for
a limited set of nanomaterials, however, an awareness of the factors affecting fate will
guide the design of field-relevant toxicity testing.
11.3
Possible Approaches to Environmental Hazard Ranking
of Nanomaterials
As has been discussed, there are currently too few data available to set appropriate
limits or environmental guideline threshold numbers that accommodate all aspects of
the various formulations, and their fate in the environment, although an approach has
been outlined that might allow these to be achieved. It might then be appropriate to
develop a basic matrix approach to assess and rank the potential environmental mobility
and hazard from nanomaterials, where bulk size, solubility and surface properties are
considered. As yet, there are insufficient data to develop such an approach.
The prediction of behaviour based on the physical or chemical properties of
nanomaterials is only possible in a very limited way. For example, the use of
quantitative structure activity relationships (QSARs) may be useful for a limited number
of comparable nanomaterial types whose structures differ in the nature of chemical
substitution on a base molecular structure (e.g. fullerenes or CNTs). Nothing has yet
been published in this area, although the Joint Research Centre (JRC) of the European
Commission in late 2006 funded the Computational Toxicology Group, within the
European Chemicals Bureau in Ispra, Italy, to review the applicability of (Q)SARs to
nanoparticles (http://ecb/jrc.it/QSAR/). While no publications have yet appeared, the
JRC website indicates that their activities are focused on the development and
harmonisation of methods for toxicity testing of nanomaterials, the in vitro test of a
representative set of manufactured nanomaterials on critical cell lines and encompass
related studies on nanometrology and reference materials as well as the development of
databases and studies on the applicability of 'in silico' methods adapting the traditional
QSAR paradigm.
A JRC report (Dearden and Worth, 2007) outlines the concept of QSPRs (quantitative
structure property relationships) as a cost-effective computational alternative to
measurement of fate and toxicity. This is being explored in relation to nanomaterials. It
is important that these models go as far as predicting actual fate rather than stopping
only with the raw material and not its in-field characteristics.
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12. RESEARCH NEEDS
It is clear that there is a need for more research to improve nanoparticle risk assessment.
This has already been discussed in a number of publications and discussed in this report.
Our recommendations are as follows:
1. There is a need for measurements in natural water, sediment and soil samples of
the stability, and short- and long-term fate of the various likely formulations that
might reach these compartments of the environment, and the development of
techniques to distinguish natural from manufactured nanoparticles. These
measurements should focus on particle concentration, size and surface
characteristics (area and charge).
2. Toxicity testing needs to be undertaken on nanoparticle formulations assessed in
(1) above. The tests should involve at least five species from four trophic levels
as required to derive PNECs using species sensitivity distributions. It is critical
that appropriate verification of particle and solute dose be undertaken in all
ecotoxicity testing, necessitating significant effort in (1) above.
3. As a precursor to toxicity testing, it will be necessary to develop standard (and
valid) methodologies for the hazard ranking of nanomaterial toxicity. These will
need to ensure the stability of the nanoparticle suspensions over the duration of
the standardised toxicity tests.
4. Comparisons of toxicity testing in natural vs. synthetic water and soil samples
demonstrating the effects of natural colloids.
13. ACKNOWLEDGEMENTS
The authors acknowledge Drs Natasha Franklin, Nicola Rogers and Simon Apte for
useful discussions and information provided for this report. We are grateful to Dr Glen
Walker (DEWHA) for his careful and comprehensive refereeing. The project was
commissioned by DEWHA with funding received from the Department of Innovation,
Industry, Science and Research under the National Nanotechnology Strategy. A 50% inkind contribution was provided by CSIRO’s Niche Manufacturing Flagship.
.
14. REFERENCES
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solubility of buckminsterfullerene in organic solvents; estimation of some physicochemical
properties. J. Chem. Soc., Perkin Trans., 2, 281-286.
Adams, L.K., Lyon, D.Y., and Alvarez, P.J.J. (2006). Comparative ecotoxicity of nanoscale
TiO2, SiO2 and ZnO water suspensions. Water Res., 40, 3527-3532.
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ANZECC/ARMCANZ (2000). Australian and New Zealand guidelines for fresh and marine
water quality. Australian and New Zealand Environment and Conservation
Council/Agriculture and Resource management Council of Australia and New Zealand,
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15. GLOSSARY
Aerobic: In the presence of oxygen
AFM: Atomic force microscopy
Agglomerate: An assemblage of particles that are rigidly bound by sintering or growth
Aggregate: An assemblage of particles that is loosely bound and are readily dispersed
AICS: Australian Inventory of Chemical Substances
Anaerobic: In the absence of oxygen
APVMA: Australian Pesticides and Veterinary Medicines Authority
Bioavailability: Available for uptake by biological organisms
n-C60: Definition of n-C60 here
CAS: Chemical Abstracts service
CNT: Carbon nanotubes
Colloid: A particle, which may be a molecular aggregate, with a diameter of 1 nm-1 µm
Cytoplasm: All of the substance of a cell outside of the nucleus
Dendrimer: A synthetic, three-dimensional molecule with branching parts, formed using a
nanoscale, multistep fabrication process. Each step results in a new “generation” that has twice
the complexity of the previous generation
DLS: Dynamic light scattering
EDX: Energy dispersive x-ray fluorescence
EM: Electron microscopy
Endocytosis: A process of cellular ingestion by which the plasma membrane folds inward to
bring substances into the cell.
Eukaryote: A single-celled or multicellular organism whose cells contain a distinct membranebound nucleus.
FFF: Field flow fractionation
Fibril: A threadlike fibre or filament
FlFFF: Flow field flow fractionation
Genotoxicity: Toxicity altering the structure or function of genetic material in an organism
Hazard quotient: The ratio of PEC to PNEC
Hydrolysis: Decomposition by reaction with water
Hydrophilic: Dissolving in or having a high affinity for water
Hydrophobic: Repelling or not easily dissolving in water
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ICPMS: Inductive coupled plasma mass spectrometry
Immunotoxicity: Toxicity affecting the functioning of the immune system
ISO: International Standards Organisation
Lipophilic: Capable of combining with or dissolving in lipids
Manufactured nanoparticles: Particles with at least one dimension smaller than 100 nm that
have been created due to deliberate human activity.
Microemulsion: An emulsion (dispersion of one immiscible liquid in another) where the
particles in the dispersed phase are less than 1000 nm, which is thermodynamically stable
MWCNT: Multi-walled carbon nanotubes
Nano: A prefix meaning one billionth (1/1,000,000,000).
Nanoclay: Naturally occurring plate-like clays with nanoparticle sizes
Nanoemulsion: An emulsion (dispersion of one immiscible liquid in another) where the
particles in the dispersed phase are less than 1000 nm. Nanoemulsions are kinetically but not
thermodynamically stable.
Nanomaterials: Materials having structured components that have one dimension lower than
100 nm. Nanoparticle threshold size can also be defined as the size leading to different physicochemical behaviours and properties than bulk material. They can be subdivided into
nanoparticles, nanofilms and nanocomposites.
Nanoparticle: Individual pieces of matter with one dimension lower than 100 nm.
Nanotechnology: Areas of technology where dimensions and tolerances in the range of 0.1 nm
to 100 nm play a critical role.
Nanotube: A one-dimensional fullerene (a convex cage of atoms with only hexagonal and/or
pentagonal faces) with a cylindrical shape.
Nanowires: One-dimensional structures, with unique electrical and optical properties, that are
used as building blocks in nanoscale devices.
NICNAS: National Industrial Chemicals Notification and Assessment Scheme
NNI: National Nanotechnology Initiative in the US
OECD: Organisation for Economic Cooperation and Development
Organelle: A differentiated structure within a cell, such as a mitochondrion, vacuole, or
chloroplast, that performs a specific function
PEC: Predicted environmental concentration
Photolysis: Chemical decomposition induced by light
PNEC: Predicted no effects concentration
Quantum dot: A nano-scale crystalline structure that can transform the colour of light. The
quantum dot is considered to have greater flexibility than other fluorescent materials, which
makes it suited to use in building nano-scale computing applications where light is used to
process information. They are made from a variety of different compounds, such as cadmium
selenide.
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ROS: Reactive oxygen species
Sonication: Treatment by high frequency sound waves
SWCNT: Single walled carbon nanotubes
TEM: Transmission electron microscopy
THF: Tetrahydrofuran
USEPA: United States Environmental Protection Agency
Zeta potential: The electrostatic potential between particles and a liquid
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