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L ` Fate of Manufactured Nanomaterials in the Australian Environment G.E. Batley and M.J. McLaughlin April 2010 (First submitted to DEWHA, September 2008) Prepared for the Department of the Environment, Water, Heritage and the Arts [Month/Year (delete if not required)]

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Fate of Manufactured Nanomaterials in the

Australian Environment G.E. Batley and M.J. McLaughlin April 2010 (First submitted to DEWHA, September 2008) Prepared for the Department of the Environment, Water, Heritage and the Arts

[Month/Year (delete if not required)]

Fate of Manufactured Nanomaterials in the Australian Environment ii

Enquiries should be addressed to:

Graeme Batley Centre for Environmental Contaminants Research CSIRO Land and Water Private Mailbag 7, Bangor NSW 2234 Phone 02 9710 6830 Fax 02 9719 6837 Email [email protected]

© Commonwealth of Australia 2010 This work is copyright. Apart from any use as permitted under the Copyright Act 1968, no part may be reproduced by any process without prior written permission from the Commonwealth. Requests and inquiries concerning reproduction and rights should be addressed to the Commonwealth Copyright Administration, Attorney General’s Department, Robert Garran Offices, National Circuit, Barton ACT 2600 or posted at http://www.ag.gov.au/cca

The views and opinions expressed in this publication are those of the authors and do not necessarily reflect those of the Australian Government or the Minister for the Environment, Heritage and the Arts or the Minister for Climate Change and Water. While reasonable efforts have been made to ensure that the contents of this publication are factually correct, the Commonwealth does not accept responsibility for the accuracy or completeness of the contents, and shall not be liable for any loss or damage that may be occasioned directly or indirectly through the use of, or reliance on, the contents of this publication.

Fate of Manufactured Nanomaterials in the Australian Environment iii

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 cadmium-containing 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 iv

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 ensure the stability of the nanoparticle suspensions over the duration of the standardised toxicity tests.

Fate of Manufactured Nanomaterials in the Australian Environment v

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 vi

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 6.2.1 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

6.3 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 .............................................. 23

7. ECOLOGICAL RISK ASSESSMENT OF MANUFACTURED NANOPARTICLES............................................................................................................................ 24

7.1 Polymeric Nanoparticles as a Separate Class .......................................................26

8. EXPOSURE ASSESSMENT .............................................................................. 26

8.1 What to Measure.....................................................................................................26

8.2 Methods for Measurement of Nanoparticles...........................................................27 8.2.1 Relevance of OECD Test Guidelines .................................................................. 28

8.3 Modelling Exposure ................................................................................................29

9. 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 Environment vii

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

10. 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 Australian Environment viii

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)

Table 8. Approach to toxicity testing of nanomaterials in waters

Fate of Manufactured Nanomaterials in the Australian Environment

ix

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

1

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 2

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 Component Use

Metal oxides Zinc oxide Cosmetic sunscreens and UV coatings; paints, plastics and packaging

Titanium dioxide Cosmetics

Cerium dioxide Automobile catalyst

Mixed oxides Cosmetics

Carbon products

Fullerenes

Single-walled and multi-walled carbon nanotubes

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

Amorphous carbon

Inks, photocopier toner, automobile tyres

Metals 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 Fe-Pd, 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

Dendrimers Coloured glasses, chemical sensors, and modified electrodes; in medicine as DNA transfecting agents,

Fate of Manufactured Nanomaterials in the Australian Environment

3

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 4

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 self-assembly 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.f-carbon.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.

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

Fate of Manufactured Nanomaterials in the Australian Environment

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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 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

Fate of Manufactured Nanomaterials in the Australian Environment 6

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)

Among the most outstanding applications of dendrimers are the formation of nanotubes, micro and macrocapsules, nanolatex, coloured glasses, chemical sensors, and modified

Fate of Manufactured Nanomaterials in the Australian Environment

7

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 published on the NICNAS website (www.nicnas.com.au) in January 2007 and are summarised in Table 2.

Fate of Manufactured Nanomaterials in the Australian Environment 8

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

Cosmetics

1-5

<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

Water treatment

10-50

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

Domestic products

Cosmetics

5-10

1-5

1-5

Zinc oxide

Surface coatings

Cosmetics

5-10

1-5

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 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

Fate of Manufactured Nanomaterials in the Australian Environment

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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 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

Fate of Manufactured Nanomaterials in the Australian Environment 10

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).

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

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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 sub-micron 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, well-mixed 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, reduced aeration (due to poor drainage), etc. These conditions are likely to act similarly on manufactured nanoparticles, although this needs confirmation.

Fate of Manufactured Nanomaterials in the Australian Environment 12

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. Nanoparticle filtration requires nanofilters, which are available, however, appropriate monitoring will be necessary to ensure their effectiveness.

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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.

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.

Atmospheric deposition

Surface runoff

Effluent discharge

Groundwater discharge Road runoff

Accidental spillage

Soil application

Fate of Manufactured Nanomaterials in the Australian Environment 14

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. FATE OF NANOMATERIALS IN AQUATIC SYSTEMS

6.1 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.

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

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of nanomaterial fate. The interactions of nanomaterials with natural colloids will play a critical role in their fate.

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 aggregation can occur through surface modification by surfactants or bulky polymeric additives.

Binding to NOM and other colloids

Aggregation

Sedimentation

Binding to NOM and other colloids

Mn+

Dissolution

Mn+

Mn+

Surfactant-stabilised nanoparticles

Binding to suspended particles/biota

Biological degradation, photolysis, hydrolysis

Fate of Manufactured Nanomaterials in the Australian Environment 16

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 THF-based 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

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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.

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.

Fate of Manufactured Nanomaterials in the Australian Environment 18

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 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

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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 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

Fate of Manufactured Nanomaterials in the Australian Environment 20

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 metal-free 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 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).

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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.

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.

Fate of Manufactured Nanomaterials in the Australian Environment 22

MNPs

1

23

4

5

1. Dissolution2. Sorption/aggregation3. Plant bioaccumulation4. Invertebrate accumulation and toxicity5. Microbial toxicity6. Direct particle uptake/toxicity7. Particle migration

6

7

Dissolvedpool

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.

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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 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

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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.

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:

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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 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.

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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. EXPOSURE ASSESSMENT

8.1 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 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

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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 to partition particles on the basis of their diameter (Giddings, 2003). For metal-containing 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

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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.

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)

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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;

(iii) runoff from contaminated soils;

(iv) aerial deposition; and

(v) emissions from wastewater treatment plants.

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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.

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

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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

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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)

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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 nano-TiO2 CNT

Compartment RE HE RE HE RE HE

Water 0.0008 0.002 >0.7 >16 0.005 0.008

Water affected by wastewater

0.2 0.5 >180 > 3900 naa na

a Not available

9. ECOTOXICOLOGY OF NANOPARTICLES

9.1 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

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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 100-500 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.

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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 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).

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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 lipid-coated 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., 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

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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.

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.

Fate of Manufactured Nanomaterials in the Australian Environment 38

Fate of Manufactured Nanomaterials in the Australian Environment

39

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-

solubilised Nominally 10-200 Standard USEPA medium Water flea Daphnia magna 48-h LC50 >35 mg/L

Zhu et al., 2006

n-C60 THF-

extract 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-

solubilised Average diameter

30

Moderately hard freshwater

USEPA protocol Water flea Daphnia magna 48-h LC50 7.9 mg/L Lovern and Klaper

(2006)

n-C60 THF

extract 10-20 Moderately hard freshwater

USEPA protocol Water flea Daphnia magna 48-h LC50 0.46 mg/L; NOEC

180 µg/L Lovern and Klaper

(2006)

n-C60 water-

solubilised, Nominal 10-200 Synthetic hard water Water flea Daphnia magna 40% mortality at 2.5 mg/L over

21 days. No acute toxicity up

to 35 mg/L

Oberdorster et al., 2006

n-C60 THF

extract Nominally 10-200 Standard USEPA medium Fathead minnow Pimephales promelas 0.5 mg/L 100% mortality in 6-

18 h

Zhu et al., 2006

n-C60 water-

solubilised, Nominally 10-200 Standard USEPA medium Fathead minnow Pimephales promelas 0.5 mg/L no effects after 48 h

Zhu et al., 2006

n-C60 THF

extract Nominally 30-100 Synthetic hard water Juvenile large-

mouth bass Mycropterus salmoides 0.8 mg/L 100% mortality in 6-

18 h

Oberdorster, 2004

Fate of Manufactured Nanomaterials in the Australian Environment 40

n-C60 water-

solubilised, Nominally 10-200 Synthetic hard water Freshwater

crustacea Hyalella azteca No toxicity below 7 mg/L Oberdorster et al., 2006

n-C60 water-

solubilised,

Nominally 10-200 Synthetic hard water Japanese medaka Oryzias latipes No acute toxicity at 0.5 mg/L

for 96h

Oberdorster et al., 2006

n-C60 THF

extract 100 nm

aggregates Synthetic hard water Zebrafish embryos Danio rerio 1.5 mg/L was toxic Zhu et al., 2007

SWCNT purified ? Seawater Meiobenthic copepods

Amphiascus tenuiremis No effects at 10 mg/L. Evidence of ingestion and aggregation

Templeton et al., 2006

SWCNT as prepared

? Seawater Meiobenthic copepods

Amphiascus tenuiremis No effect at 1.6 mg/L; effects at 10 mg/L.

Templeton et al., 2006

SWCNT ? Freshwater with up to 0.15 mg/L SDS

Rainbow trout Oncorhynchus mykiss Effects on ventilation rate, gill pathologies and gill mucus secretion at 0.5 mg/L

Smith et al., 2007

SWCNT ? Freshwater and seawater Zebrafish embryos Danio rerio Hatching delay at 150 mg/L Cheng et al., 2007

TiO2 Nominal 25 (small); 100 (large)

Moderately hard water OECD 201 protocol

Algae Desmodesmus subspicatus

Chlorophyll fluorescence 72-h EC50 44 mg/L small; no dose response large

Hund-Rinke and Simon, 2006

TiO2 Average 140 Moderately hard water

OECD 201 protocol

Algae Pseudokirchneriella subcapitata

Chlorophyll fluorescence

72h EC50 16-21 mg/L Warheit et al., 2007

TiO2 Nominal 25 (small); 100 (large)

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

TiO2 THF dispersed; sonicated

30 THF; 100-500 sonicated

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

TiO2 THF dispersed; sonicated

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

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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 No mortality during 14-day exposure up to 1.0 mg/L. Sub-lethal effects including gill damage, observed.

Federici et al., 2007

TiO2 140

Moderately hard water OECD 201 protocol

Rainbow trout Oncorhynchus mykiss 96-h EC50 >100 mg/L

Warheit et al., 2007

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 No mortality with 25 day exposure to 10 mg/L TiO2. Increased Cd accumulation

Zhang et al., 2007

ZnO Average 178-361 USEPA, pH 7.5 Algae Pseudokirchneriella

subcapitata

72-h EC50 68µg/L due to dissolved Zn

Franklin et al., 2007

ZnO Mixed Spring water + food pellets Water flea Daphnia magna 8-day EC50 0.2-0.5 mg/L,

possibly dissolved Zn Adams et al., 2006

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

10-20 ng/L

Lee et al., 2007

Quantum dots;

Cd/Se or Cd/Te

core with ZnS

Estimated 10-25 Moderately hard water Water flea Ceriodaphnia dubia 96-h LC50 >110 µg/L Bouldin et al., 2008

Fate of Manufactured Nanomaterials in the Australian Environment 42

shell

Quantum dots;

Cd/Se or Cd/Te

core with ZnS

shell

Estimated 10-25 Moderately hard water Algae Pseudokirchneriella subcapitata

96-h LC50 37.1 µg/L of

quantum dots, estimated as 9.6

µg/L Cd and 2.4 µg/L Se

Bouldin et al., 2008

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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 Estimated chronic NOEC,

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 16-21

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 44

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

Bacillus subtilis Whole cell TEM, fluorescence

spectroscopy show adenine- Kloepfer et al., 2005

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conjugated QDs < 5 nm are internalised. Intracellular Cd and Se confirmed.

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-coated

Daphnia magna 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

Fish

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

Gut

Histopathological changes to the gill and gut but fish did not accumulate TiO2 in the internal organs.

Federici et al., 2007

SWCNT Oncorhynchus mykiss

Gill

Gut

Histopathological changes to the gill and gut and liver. Aggregated SWCNTs observed in the gut lumen.

Smith et al., 2007

9.2.5 Ecological Impacts

There have been no published studies on the broader ecological impacts of manufactured nanoparticles.

Fate of Manufactured Nanomaterials in the Australian Environment 46

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

Carbon-containing

A) Fullerenes

C60 granular and

C60 water suspension (n-C60)

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

n-C60 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

B) Carbon nanotubes

Multi-walled No effect on seed germination and root growth of corn, cucumber, lettuce, radish, and rape. Reduced root growth of ryegrass.

Lin and Xing, 2007

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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

Metal oxides

Al2O3 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

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 48

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

Fate of Manufactured Nanomaterials in the Australian Environment 50

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 52

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 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;

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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 (n-C60 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?

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(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.

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

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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;

(vii) For metallic nanomaterials, water-soluble and potentially dissociated nanoparticles are likely to be more toxic than insoluble and undissociated

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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% in-kind contribution was provided by CSIRO’s Niche Manufacturing Flagship. .

14. REFERENCES

Abraham, M.H., Green, C.E., and Acree, W.E. (2000). Correlation and prediction of the 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.

Fate of Manufactured Nanomaterials in the Australian Environment 60

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, Canberra ACT, Australia

Apte, S.C., Rogers, N.T., and Batley, G.E. (2008). Ecotoxicology of manufactured nanomaterials. In: Environmental and human health effects of nanoparticles, Lead, J.R. (ed) in preparation.

APVMA (2008). The APVMA and nanotechnology. Australian Pharmaceutical and Veterinary Medicines Authority position paper. http://www.apvma.gov.au/new/downloads/Nanotechnology.pdf

Arabe, K.C. (2003). Nanomaterials set for explosive growth. http://news.thomasnet.com/IMT/archives/2003/09/nanomaterials_s.html

Baalousha, M., Manciulea, A., Cumberland, S., Kendall, K., and Lead, J.R. (2008). Aggregation and surface properties of iron oxide nanoparticles: influence of pH and natural organic matter. Environ.Toxicol.Chem., 27, 1875-1852.

Becker, L., Poreda, R.J., and Bada, J.L. (1996). Extraterrestrial helium trapped in fullerenes in the Sudbury Impact Structure: Science, 272, 249- 252.

Benn, T.M., and Westerhoff, P. (2008). Nanoparticle silver released into water from commercially available sock fabrics. Environ. Sci. Technol., 42, 4133-4139.

Bergeron, S., and Archambault, E. (2005). Canadian stewardship practices for environmental nanotechnology. Science-Metrix Report prepared for Environment Canada, Montreal, Canada. Blaser, S.A., Scheringer, M., MacLeod, M., and Hungerbühler, K. (2008). Estimation of cumulative aquatic exposure and risk due to silver: Contribution of nano-functionalized plastics and textiles. Sci. Total. Environ., 390, 396-409. Bootz, A., Vogel, V., Schubert, D., and Kreuter, J. (2004). Comparison of scanning electron microscopy, dynamic light scattering and analytical ultracentrifugation for the sizing of poly(butylcyanoacrylate) nanoparticles. Europ. J. Pharmaceut. Biopharmaceut., 57, 369–375. Borm, P., Klaessig, F.C., Landry, T.D., Moudgil, B., Pauluhn, J., Thomas, K., Trottier, R., and Wood, S. (2006). Research strategies for safety evaluation of nanomaterials, Part V: role of dissolution in biological fate and effects of nanoscale particles. Toxicol. Sci., 90, 23–32. Bouldin, J.L., Ingle, T.M., Sengupta, A., Alexander, R., Hannigan, R.E., and Buchanan, R.A. (2008). Aqueous toxicity and food chain transfer of quantum dots in freshwater algae and Ceriodaphnia dubia. Environ. Toxicol. Chem., 27, 1958-1963. Boxall, A.B.A., Chaudhry, Q., Sinclair, C., Jones, A., Aitken, R., Jefferson, B., and Watt, C. (2007). Current and future predicted environmental exposure to engineered nanoparticles. Central Science Laboratory Report, University of York, prepared for the UK Department of Environment, Food and Rural Affairs.

Fate of Manufactured Nanomaterials in the Australian Environment

61

Brant, J., Lecoanet, H. and Wiesner, M.R. (2005a). Aggregation and deposition characteristics of fullerene nanoparticles in aqueous solution. J. Nanoparticle Res., 7, 545-553. Brant, J., Lecoanet, H., Hotze, M., and Wiesner, M. (2005b). Comparison of electrokinetic properties of colloidal fullerenes (n-C60) formed using two procedures. Environ. Sci. Technol., 39, 6343-6351. Brayner, R., Ferrari-Iliou, R., Brivois, N., Djediat, S., Benedetti, M.F., and Fievet, F. (2006). Toxicological impact studies based on Escherichia coli bacteria in ultrafine ZnO nanoparticles colloidal medium. Nano Lett., 6, 866-870. Buffle, J., and Leppard, G.G. (1995a). Characterisation of aquatic colloids and macromolecules. 1. Structure and behaviour of colloidal material. Environ. Sci. Technol., 29, 2169-2175. Buffle, J., and Leppard, G.G. (1995b). Characterisation of aquatic colloids and macromolecules. 2. Key role of physical structures on analytical results. Environ. Sci. Technol., 29, 2176-2184. Buffle, J., Wilkinson, K.J., Stoll, S., Filella, M., and Zhang, J. (1998). A generalized description of aquatic colloidal interactions; the three-colloidal component approach. Environ. Sci. Technol., 32, 2887-2899. Buzea, C., Pacheco, I.I., and Robbie, K. (2007). Nanomaterials and nanoparticles: sources and toxicity. Biointerphases, 2, 17-71. Cameron, F.K. (1915). Soil colloids and the soil solution. J. Phys. Chem., 19, 1-13.

CBEN (2005). Center for Biological and Environmental Technology, Rice University website. http://cben.rice.edu//index.cfm Chen, K.L., and Elimelech, M. (2007). Influence of humic acid on the aggregation kinetics of fullerene (C60) nanoparticles in monovalent and divalent electrolyte solutions. J. Colloid Interface Sci., 309, 126-34. Cheng, J., Flahaut, E., and Cheng, S.H. (2007). Effect of carbon nanotubes on developing zebrafish (Danio rerio) embryos. Environ. Toxicol. Chem., 26, 708-716. Choi, O., Deng, K.K., Kim, N-J., Ross, L., Surampalli, R.Y., and Hu, Z. (2008). The inhibitory effects of silver nanoparticles, silver ions, and silver chloride colloids on microbial growth. Water Res., 42, 3066-3074. Choy, J.H., Kwak, S.Y., Jeong, Y.J., and Park, J.S. (2000). Inorganic layered double hydroxides as nonviral vectors. Angew. Chem. Int. Ed., 39, 4042-4044. Crane, M., and Handy, R. (2007). An assessment of regulatory testing strategies and methods for characterising the ecotoxicological hazards of nanomaterials. Watts and Crane Associates Report prepared for DEFRA, 147 pp.

Fate of Manufactured Nanomaterials in the Australian Environment 62

Dearden, J., and Worth, A. (2007). In silico prediction of physicochemical properties. Joint Research Centre, European Commission Scientific and Technical Report EUR 23051 EN – 2007, 68 pp. DEFRA (2005). Characterising the potential risks posed by engineered nanoparticles. UK Department for Environment, Food and Rural Affairs Report (http://www.defra.gov.uk) DEFRA (2007). Characterising the potential risks posed by engineered nanoparticles. UK Department for Environment, Food and Rural Affairs Report (http://www.defra.gov.uk) Degueldre, C., Favarger, P-Y., and Wold, S. (2006). Gold colloid analysis by inductively coupled plasma-mass spectrometry in a single particle mode. Anal. Chim. Acta, 555, 263–268 Derfus, A.M., Chan, W.C.W., and Bhatia, S.N. (2004). Probing the nanotoxicity of semiconductor quantum dots. Nano Lett., 4, 11-18. DEW (2007). Environmental risk assessment guidance manual for industrial chemicals. Environment Protection Branch, Department of the Environment and Water Resources Report, Canberra, ACT, Australia. Doty, R.C., Tshikhudo, T.R., and Brust, M., (2005). Extremely stable water-soluble Ag nanoparticles. Chem. Mater., 17, 4630-4635. Dowling, A. (2005). Is nanotechnology “the next GM”? New Scientist, 2490, March 12, 19. Duffin, R., Tran, C.L., Clouter, A., Brown, D.M., MacNee, W., Stone, V., and Donaldson, K. (2002). The importance of surface area and specific reactivity in the acute pulmonary inflammatory response to particles. Ann. Occup. Hyg., 46 (Suppl. 1), 242-245. Electronics.ca (2007). Current and future market for nano and synthetic clays 2007-2012. http://www.electronics.ca/reports/nanotechnology/synthetic_clays.html. Environmental Defense-DuPont (2007). Nano risk framework. (http://nanoriskframework.com/page.cfm?tagID=1081) EPHC (2006). NChEM: a national framework for chemicals management in Australia. Environment Protection and Heritage Council, Canberra ACT, Australia. Fang, J.S., Lyon, D.Y., Wiesner, M.R., Dong, J.P., and Alvarez, P.J.J. (2007). Effect of a fullerene water suspension on bacterial phospholipids and membrane phase behaviour. Environ. Sci. Technol., 41, 2636-2642. Federici, G., Shaw, B.J., and Handy, R.D. (2007). Toxicity of titanium dioxide nanoparticles to rainbow trout (Oncorhynchus mykiss): Gill injury, oxidative stress, and other physiological effects. Aquat. Toxicol., 84, 415-430. Fernandes, M., Rosenkranz, P., Ford, A., Christofi, N., and Stone V. (2006). Ecotoxicology of nanoparticles (NPs). SETAC Globe, September, 43-45.

Fate of Manufactured Nanomaterials in the Australian Environment

63

Fortner, J.D., Lyon, D.Y., Sayes, C.M., Boyd, A.M., Falkner, J.C., Hotze, E.M., Alemany, L.B., Tao, Y.J., Guo, W., Ausman, K.D., Colvin, V.L., and Hughes, J.B. (2005). C60 in water: Nanocrystal formation and microbial response. Environ. Sci. Technol., 39, 4307-4316.

Franklin, N.M., Rogers, N.T., Apte, S.C., Batley, G.E. and Casey, P.E. (2007). Comparative toxicity of nanoparticulate ZnO, bulk ZnO and ZnCl2 to a freshwater microalga (Pseudokirchnerilla subcapitata): the importance of particle solubility. Environ. Sci. Technol., 41, 8484-8490. Frater, L., Stokes, E., Lee, R., and Oriola, T. (2006). An overview of the framework of current regulation affecting the development and marketing of nanomaterials. A Report for the DTI. ESRC Centre for Business Relationships Accountability Sustainability and Society (BRASS). Cardiff University. (http://www.dti.gov.uk/science/science-in-govt/st_policy_issues/nanotechnology/page20218.html) Frechet J.M.J., and Tomalia. D.A. (2002). Dendrimers and other Dendritic Polymers, John Wiley and Sons, NY, NY, USA. Ghafari, P., St-Denis, C.H., Power, M.E., Jin, X., Tsou, V., Mandal, H.S., Bols, N.C., and Tang, X. (2008). Impact of carbon nanotubes on the ingestion and digestion of bacteria by ciliated protozoa. Nature Nanotech., 3, 347-351.

Giddings, J.C. (1993). Field-flow fractionation: analysis of macromolecular, colloidal, and particulate materials. Science, 260, 1456-1465. Greenwood, M. (2007). Thinking big about things small: creating an effective oversight system for nanotechnology. Woodrow Wilson International Center for Scholars Project on Emerging Nanotechnologies Report, Washington, DC, USA. Griffitt, R.J., Weil, R., Hyndman, K.A., Denslow, N.D., Powers, K., Taylor, D. and Barber, D.S. (2007). Exposure to copper nanoparticles causes gill injury and acute lethality in zebrafish (Danio rerio). Environ. Sci. Technol., 41, 8178–8186. Griffitt, R.J., Luo, J., Gao, J., Bonzongo, J-C., Barber, D.S. (2008). Effects of particle composition and species on toxicity of metallic nanomaterials to aquatic organisms. Environ. Toxicol. Chem., 27, 1972-1978. Gustafsson, O., and Gschwend, P.M. (1997). Aquatic colloids: concepts, definitions and current challenges. Limnol. Oceanogr., 42, 519-528. Haddon, R.C., Sippel, J., Rinzler, A.G., and Papdimitrakopoulos, F. (2004). Purification and separation of carbon nanotubes. MRS Bull., 29, 252–9. Handy, R.D., Von der Kammer, F., Lead, J.R., Hassessov, M., Owe, R., and Crane, M. (2008). The ecotoxicology and chemistry of manufactured nanoparticles. Ecotox., 17, 287-314.

Fate of Manufactured Nanomaterials in the Australian Environment 64

Haruta, M., Yamada, N., Kobayashi, T., and Iijima, S. (1989). Gold catalysts prepared by coprecipitation for low-temperature oxidation of hydrogen and of carbon monoxide. J. Catal., 115, 301-309. Hassellov, M., Readman, J.W., Ranville, J.F., and Tiede, K. (2008). Nanoparticle analysis and characterisation methodologies in environmental risk assessment of engineered nanoparticles. Ecotox., 17, 344-361. Holzinger, M., Steinmetz, J., Samaille, D., Glerup, M., Paillet, M., Bernier, P., Ley, L., and Graupner. R. (2004). [2+1] cycloaddition for cross-linking SWCNTs. Carbon, 42, 941-947. Hu, X., Liu, J., Mayer, P., and Jiang, G. (2008). Impacts of some enviromentally relevant parameters on the sorption of polycyclic aromatic hydrocarbons to aqueous suspensions of fullerene. Environ. Toxicol. Chem., 27, 1868-1874. Hund-Rinke, K., and Simon, M. (2006). Ecotoxic effect of photocatalytic active nanoparticles (TiO2) on algae and daphnids. Environ. Sci. Pollut. Res., 13, 225-232. Hydutsky, B.W., Mack, E.J., Beckerman, B.B., Skluzacek, J.M., and Mallouk, T.E. (2007). Optimization of nano- and microiron transport through sand columns using polyelectrolyte mixtures. Environ. Sci. Technol., 41, 6418-6424.

Hyung, H., Fortner, J.D., Hughes, J.B., and Kim, J-H. (2007). Natural organic matter stabilization of multi-walled nanotubes in water. Environ. Sci. Technol., 41, 179-184. Inoue, K., Takano, H., Sakurai, M., Oda, T., Tamura, H., Yanagisawa, R., Shimada, A., and Yoshikawa, T. (2006). The role of toll-like receptor 4 in airway inflammation induced by diesel exhaust particles. Arch. Toxicol., 80, 275-279.

Jemec, A., Drobne, D., Remskar, M., Sepcic, K. and Tisler, T. (2008). Effects of ingested nano-sized titanium dioxide on terrestrial isopods (Porcellio scaber). Environ. Toxicol. Chem., 27, 1904-1914.

Johansen, A., Pedersen, A.L., Karlson, U., Hansen, B.J., Scott-Fordsmand, J.J., and Winding, A. (2008). Effects of C60 fullerene nanoparticles on soil bacteria and protozoans. Environ. Toxicol. Chem., 27, 1895-1903.

Kang, S., Pinault, M., Pfefferle, L.D., and Elimelech, M. (2007). Single-walled carbon nanotubes exhibit strong antimicrobial activity. Langmuir, 23, 8670-8673.

Kaplan, D.I., Sumner, M.E., Bertsch, P.M., and Adriano, D.C. (1996). Chemical conditions conducive to the release of mobile colloids from Ultisol profiles. Soil Science Soc. Amer. J., 60, 269-274.

Ke, P.C., and Qiao, R. (2007). Carbon nanomaterials in biological systems. J. Phys. Condens. Matt., 19, 373101.

Fate of Manufactured Nanomaterials in the Australian Environment

65

Kennedy, A.J., Hull, M.S., Steevens, J.A., Dontsova, K.M., Chappell, M.A., Gunter, J.C., and Weiss, C.A. (2008). Factors influencing the partitioning and toxicity of nanotubes in the aquatic environment. Environ. Toxicol. Chem., 27, 1932-1941.

Kirchner, C., Liedl, T., Kudera, S., Pellegrino, T., Javier, A.M., Gaub, H.E., Stolzle, S., Fertig, N., and Park, W.J. (2005). Cytotoxicity of colloidal CdSe and CdSe/ZnS nanoparticles. Nano Lett., 5, 331-338. Klaine, S.J., Alvarez, P.J.J., Batley, G.E., Fernandes, T.F., Handy, R.D., Lyon, D., Mahendra, S., McLaughlin, M.J., and Lead, J.R. (2008). Nanomaterials in the environment: behaviour, fate, bioavailability and effects. Environ. Toxicol. Chem., 27, 1825-1851. Kloepfer, J.A., Mielke, R.E., and Nadeau, J.L. (2005). Uptake of CdSe and CdSe/ZnS quantum dots into bacteria via purine-dependent mechanisms. App. Environ. Microbiol., 71, 2548-2557. Kramer, J.R., Benoit, G., Bowles, K.C., Di Toro, D.M., Herrin, R.T., Luther, G.W., Manalopoulis, H., Robilliard, K.A., Shafer, M.M., and Shaw, J.R. (2002). Environmental chemistry of silver. In: Silver in the Environment: Transport, Fate, and Effects. Andren, A.W. and Bober, T.W. (eds), SETAC Press, Pensacola, FL, USA, pp. 1-25 Kretzschmar, R., and Schafer, T. (2005). Metal retention and transport on colloidal particles in the environment. Elements, 1, 205-210. Kretzschmar, R., Robarge, W.P., and Amoozegar, A. (1995). Influence of natural organic-matter on colloid transport through saprolite. Water Resourc. Res., 31, 435-445.

Kroto, H.W., and Walton, D.R.M. (2007). In Encyclopædia Britannica. http://www.britannica.com/eb/article-234438 Kwak, S.Y., Kim, S.H., and Kim, S.S. (2001). Hybrid organic/inorganic reverse osmosis (RO) membrane for bactericidal anti-fouling. 1. Preparation and characterization of TiO2 nanoparticle self-assembled aromatic polyamide thin-film-composite (TFC) membrane. Environ. Sci. Technol., 35, 2388-2394. Lawrence, J.J., and Warisnoicharoen, W. (2006). Recent advances in microemulsions as drug delivery vehicles. In: Nanoparticulates as Drug Carrier, Torchilin, V.P. (ed), Imperial College Press, London, UK, pp. 125-172. Lead, J.R., and Wilkinson, K.J. (2007). Environmental colloids and particles: Current knowledge and future developments. In: Environmental Colloids and Particles: Behaviour, Separation and Characterisation, Wilkinson, K.J., and Lead, J.R. (eds), John Wiley and Sons, Chichester, UK., pp 1-16. Lecoanet, H.F., Bottero, J.Y., and Wiesner, M.R. (2004). Laboratory assessment of the mobility of nanomaterials in porous media. Environ. Sci. Technol., 38, 5164–5169. Lee, K.J., Nallathamby, P.D., Browning, L.M., Xu, X.H., and Osgood, C.C.J. (2007). In vivo imaging of transport and biocompatibility of single silver nanoparticles in early development of zebrafish embryos. ACS Nano, 1, 133–143.

Fate of Manufactured Nanomaterials in the Australian Environment 66

Lee, W.M., An, Y.J., Yoon, H. and Kweon, H.S. (2008). Toxicity and bioavailability of copper nanoparticles to the terrestrial plants mung bean (Phaseolus radiatus) and wheat (Triticum aestivum): Plant agar test for water-insoluble nanoparticles. Environ. Toxicol. Chem., 27, 1915-1921.

Lekas, D. (2005). Analysis of nanotechnology from an industrial ecology perspective Part II: Substance flow analysis of carbon nanotubes. Project on Emerging Nanotechnologies Report, Woodrow Wilson International Centre for Scholars, 22 pages. Li, D., Lyon, D.Y., Li, Q., and Alvarez, P.J.J. (2008). Effect of soil sorption and aquatic natural organic matter on the antibacterial activity of a fullerene water suspension. Environ. Toxicol. Chem., 27, 1888-1894.

Liang, L., and Morgan, J.J. (1990). Chemical aspects of iron oxide coagulation in water: Laboratory studies and implications for natural systems. Aquat. Sci., 52, 32-55.

Limbach, L.K., Bereiter, R., Muller, E., Krebs, R., Galli, R., and Stark, W.J. (2008). Removal of oxide nanoparticles in a model wastewater treatment plant: influence of agglomeration and surfactants on clearing efficiency. Environ. Sci. Technol., 42, 5828-5833.

Lin, D., and Xing, B. (2007). Phytotoxicity of nanoparticles: Inhibition of seed germination and root growth. Environ. Pollut., 150, 243-250.

Lindsay, W.L. (1979). Chemical Equilibria in Soils. John Wiley and Sons, New York, USA.

Lovern, S.B., and Klaper, R. (2006). Daphnia magna mortality when exposed to titanium dioxide and fullerene (C60) nanoparticles. Environ. Toxicol. Chem., 25:1132-1137.

Lovern, S.B., Strickler, J.R., and Klaper, R. (2007). Behavioral and physiological changes in Daphnia magna when exposed to nanoparticle suspensions (titanium dioxide, nano-C60, and C60HxC70Hx). Environ.Sci. Technol., 41, 4465-4470. Ludlow, K., Bowman, D., and Hodge, G. (2007). A review of possible impacts of nanotechnology on Australia’s regulatory framework. Monash University, Faculty of Law Report, 110 pp.

Lyon, D.Y., Adams, L.K., Falkner, J.C., and Alvarez, J.J. (2006). Antibacterial activity of fullerene water suspensions: effects of preparation method and particle size. Environ. Sci. Technol., 40, 4360-4366. Mackay, C.E., Johns, M., Salatas, J.H., Bessinger, B., and Perri, M. (2006). Stochastic probability modelling to predict the environmental stability of nanoparticles in aqueous suspension. Integ. Environ. Assess. Manage., 2, 293-298. Magistad, O.C. (1925). The aluminium content of the soil solution and its relation to soil reaction and plant growth. Soil Sci., 20, 181-211.

Fate of Manufactured Nanomaterials in the Australian Environment

67

Makhluf, S., Dror, R., Nitzan, Y., Abramovich, Y., Jelinek, R., and Gedanken, A. (2005). Microwave-assisted synthesis of nanocrystalline MgO and its use as a bacteriocide. Adv. Funct. Mater., 15, 1708-1715. Maynard, A.D. (2006). Nanotechnology: a research strategy for addressing risk. Woodrow Wilson International Center for Scholars Project on Emerging Nanotechnologies Report, Washington, DC, USA. Meyer, D.E., Wood, K., Bachas, L.G., and Battacharyya, D. (2004). Degradation of chlorinated organics by membrane-immobilzed nanosized metals. Environ. Prog. 23, 232-242. Mondal, K., Jegadeesan, G., and Lalvani, S.B. (2004). Removal of selenate by Fe and NiFe nanosized particles. Ind. Eng. Chem. Res., 43, 4922-.4935 Moore, M.N. (2006). Do nanoparticles present ecotoxicological risks for the health of the aquatic environment? Environ. Intern., 32, 967-976 Morones, J.R., Elechiguerra, J.L., Camacho, A., Holt, K., Kouri, J.B., Ramirez, J.T., and Yacaman, M.J. (2005). The bactericidal effect of silver nanoparticles. Nanotech., 16, 2346-2353. Mueller, N.C., and Nowack, B. (2008). Exposure modeling of engineered nanoparticles in the environment. Environ. Sci. Technol., 42, 4447-4453. Naidu, R. and Rengasamy, P. (1993). Ion interactions and constraints to plant nutrition in Australian sodic soils. Aust., J. Soil Res., 31, 801-819.

Nel, A., Xia, T., Madler, L., and Li, N. (2006). Toxic potential of materials at the nanolevel; Science, 311, 622-627. Nichols, G., Byard, S., Bloxham, M.J., Botterill, J., Dawson, N.J., Dennis, A., Diart, V., North, N.C., and Sherwood, J.D. (2002). A review of the terms agglomerate and aggregate with a recommendation for nomenclature used in powder and particle characterisation. J. Pharmaceut. Sci., 91, 2103-2109. NICNAS (2000). Acrylic Latex 99 R 9502, File No. PLC/145, National Industrial Chemicals Notification and Assessment Scheme, Sydney, Australia.

NNI (2005). National Nanotechnology Initiatives website. http://www.nano.gov NNST (2006). Options for a National Nanotechnology Strategy, National Nanotechnology Strategy Taskforce available at: http://www.innovation.gov.au/Section/Innovation/Pages/OptionsforaNationalNanotechnologyStrategyReport.aspx. Noack, A.G., Grant, C.D., and Chittleborough, D.J. (2000). Colloid movement through stable soils of low cation-exchange capacity. Environ. Sci. Technol., 34, 2490-2497.

Nowack, B., and Buschelli, T.D. (2007). Occurrence, behaviour and effects of nanoparticles in the environment. Environ. Pollut., 150, 5-22

Fate of Manufactured Nanomaterials in the Australian Environment 68

Nyberg, L., Turco, R.F., and Nies, L. (2008) Assessing the impact of nanomaterials on anaerobic microbial communities. Environ. Sci. Technol., 42, 1938-1943.

Oberdorster, E. (2004). Manufactured nanoparticles (fullerenes, C60) induce oxidative stress in the brain of juvenile largemouth bass. Environ. Health Perspect., 112, 1058-1062. Oberdorster, E., Zhu, S., Blickley, T.M., McClellan-Green, P., and Haasch, M.L. (2006). Ecotoxicology of carbon-based engineered NPs: Effects of fullerene (C60) on aquatic organisms. Carbon, 44, 1112-1120. Oberdorster, G., Oberdörster, E., and Oberdörster, J. (2005). Nanotoxicology: an emerging discipline evolving from studies of ultrafine particles. Environ. Health Perspect., 113, 823–839. OECD (2007). OECD Guidelines for the Testing of Chemicals. Organisation for Economic Cooperation and Development (http://www.oecd.org/document/40/0,3343,en_2649_34377_37051368_1_1_1_1,00.html) Owen, R., and Handy, R., (2007). Formulating the problems for environmental risk assessment of nanoparticles. Environ. Sci. Technol., 41, 5582-5588. PEN (2007a). Woodrow Wilson International Center for Scholars Project on Emerging Nanotechnologies website: http://www.nanotechproject.org PEN (2007b). Consumer Product Inventory. Woodrow Wilson International Centre for Scholars Project on Emerging Nanotechnologies, Washington DC, USA. Petersen, E.J., Huang, Q.G., and Weber, W.J. (2008). Bioaccumulation of radio-labeled carbon nanotubes by Eisenia foetida. Environ. Sci. Technol. 42, 3090-3095.

Phenrat, T., Saleh, N., Sirk, K., Tilton, R.D., and Lowry, G.V. (2007). Aggregation and sedimentation of aqueous nanoscale zerovalent iron dispersions. Environ. Sci. Technol., 41, 284-290. Plata, D.L., Gschwend, P.M., and Reddy, C.M. (2008). Industrially synthesized single-walled carbon nanotubes: compositional data for users, environmental risk assessments, and source apportionment. Nanotech., 19, 185706. Pulskamp, K., Diabaté, S., and Krug, H.F. (2005). Carbon nanotubes show no sign of acute toxicity but induce intracellular reactive oxygen species in dependence on contaminants. Toxicol. Lett., 168, 58-74. Qiao, R., and Ke, P.C. (2006). Lipid-carbon nanotube self assembly in aqueous solution. J. Am. Chem. Soc., 128, 13656-13657. Raja, R., Sankar, G., Hermans, S., Shephard, D.S., Bromley, S., Thomas, J.M., and, Johnson, B.F.G. (1999). Preparation and characterisation of a highly active bimetallic (Pd–Ru) nanoparticle heterogeneous catalyst. Chem. Commun., 1571–1572.

Fate of Manufactured Nanomaterials in the Australian Environment

69

Reddy, K.M., Feris, K., Bell, J., Wingett, D.G., Hanley, C., and Punnoose, A. (2007). Selective toxicity of zinc oxide nanoparticles to prokaryotic and eukaryotic systems. App. Phys. Lett., 90, 213902. Rengasamy, P., and Olsson, K.A. (1991). Sodicity and soil structure. Aust. J. Soil Res., 29, 935-952.

Roberts, A.P., Mount, A.S., Seda, B., Souther, J., Qiao, R., Lin, S., Ke, P.C., Rao, A.M., and Klaine, S.J. (2007). In vivo biomodification of lipid-coated carbon nanotubes by Daphnia magna. Environ. Sci. Technol., 41, 3025-3029. Royal Society/Royal Academy of Engineering (2004). Nanoscience and nanotechnologies: opportunities and uncertainties. Royal Society Policy Document 19/04. Saleh, N., Kim, H.J., Phenrat, T., Matyjaszewski, K., Tilton, R.D. and Lowry, G.V. (2008). Ionic strength and composition affect the mobility of surface-modified Fe0 nanoparticles in water-saturated sand columns. Environ. Sci. Technol., 42, 3349-3355.

Sayes, C.M., Fortner, J.D., Guo, W., Lyon, D., Boyd, A.M., Ausman, K.D., Tao, Y.J., Sitharaman, B., and Wilson, L.J. (2004). The differential cytotoxicity of water-soluble fullerenes. Nano Lett., 4, 1881-1887. SCENIHR (2005). Opinion on the appropriateness of existing methodologies to assess the potential risks associated with engineered and adventitious products of nanotechnologies. Scientific Committee on Emerging and Newly Identified Health Risks Report SCENIHR/002/05, European Commission Health and Consumer Protection Directorate-General, 78 pp. Schrick, B., Hydutsky, B.W., Blough, J.L., and Mallouk, T.E. (2004). Delivery vehicles for zerovalent metal nanoparticles in soil and groundwater. Chem. Mater., 16, 2187. Seaman, J.C., and Bertsch, P.M. (2000). Selective colloid mobilization through surface-charge manipulation. Environ. Sci. Technol., 34, 3749-3755.

Seaman, J.C., Bertsch, P.M., and Strom, R.N. (1997). Characterisation of colloids mobilized from southeastern coastal plain sediments. Environ. Sci. Technol., 31, 2782-2790.

Shi, J.P., Evans, D.E., Khan, A.A., and Harrison, R.M. (2001). Sources and concentration of nanoparticles (<10 nm diameter) in the urban atmosphere. Atmos. Environ., 35, 1193-1202. Siepmann, R., Von Der Kammer, F., and Forstner, U. (2004). Colloidal transport and agglomeration in column studies for advanced run-off filtration facilities - particle size and time resolved monitoring of effluents with flow-field-flow-fractionation. Water Sci. Technol., 50, 95-102.

Smith, C.J., Shaw, B.J., and Handy, R.D. (2007). Toxicity of single walled carbon nanotubes to rainbow trout, (Oncorhynchus mykiss): Respiratory toxicity, organ pathologies, and other physiological effects. Aquat. Toxicol., 82, 94-109.

Fate of Manufactured Nanomaterials in the Australian Environment 70

Sondi, I., and Salopek-Sondi, B. (2004). Silver nanoparticles as antimicrobial agent: a case study on E. coli as a model for Gram-negative bacteria. J. Colloid Interface Sci., 275, 177-182. Stoimenov, P.K., Klinger, R.L., Marchin, G.L., and Klabunde, K.J. (2002). Metal oxide nanoparticles as bactericidal agents. Langmuir, 18, 6679-6686 Stolpe, B., Hassellöv, M., Andersson, K., and Turner, D.R. (2005). High resolution ICPMS as an on-line detector for flow field-flow fractionation; multi-element determination of colloidal size distributions in a natural water sample. Anal. Chim. Acta, 535, 109-121. Templeton, R.C., Ferguson, P.L., Washburn, K.M., Scrivens, W.A., and Chandler, G.T. (2006). Life-cycle effects of single-walled carbon nanotubes (SWNTs) on an estuarine meiobenthic copepod. Environ. Sci. Technol., 40, 7387-7393. Terashima, M., and Nagao, S. (2007). Solubilization of (60) fullerene in water by aquatic humic substances. Chem. Lett., 36, 302-303. TGA (2006). A review of the scientific literature on the safety of nanoparticulate titanium dioxide or zinc oxide in sunscreens, Therapeutic Goods Administration Report http://www.tga.gov.au/npmeds/sunscreen-zotd.htm Tiede, K., Boxall, A.B.A., Tear, S.P., Lewis, J., David, H. and Hassellov, M. (2008). Detection and characterization of engineered nanoparticles in food and the environment. Food Addit. Contam., 25, 795-821.

Tipping, E., and Higgins, D.C. (1982). The effect of adsorbed humic substances on the colloid stability of haematite particles. Coll. Surf., 5, 85–92. Tong, Z.H., Bischoff, M., Nies, L., Applegate, B., and Turco, R.F. (2007). Impact of fullerene (C-60) on a soil microbial community. Environ. Sci. Technol., 41, 2985-2991. Tratnyek, P.G., and Johnson, R.L. (2006). Nanotechnologies for environmental cleanup. NanoToday, 1, 44-48. Turkevich, J., Garton, G., and Stevenson, P.C. (1954). The color of colloidal gold. J. Colloid Sci., 9, 26-35. USDOE (2007). Approach to Nanomaterial ES and H. Nanoscale Science Research Center Report, Revision 2, US Department of Energy, 23 pages. USEPA (2007). Nanotechnology white paper. US Environmental Protection Agency Report EPA 100/B-07/001, Office of the Science Advisor, Washington, DC, USA. USEPA (2008). Draft nanomaterial research strategy. US Environmental Protection Agency Report EPA/600/S-08/002, Office of Research and Development, Washington, DC, USA Wang, Y.G., Li, Y.S. and Pennell, K.D. (2008). Influence of electrolyte species and concentration on the aggregation and transport of fullerene nanoparticles in quartz sands. Environ. Toxicol. Chem., 27, 1860-1867.

Fate of Manufactured Nanomaterials in the Australian Environment

71

Warheit, D.B., Hoke, R.A., Finlay, C., Donner, E.M., Reed, K.L., and Sayes, C.M. (2007). Development of a base set of toxicity tests using ultrafine TiO2 particles as a component of nanoparticle risk management. Toxicol. Lett., 171, 99-110. Wiesner, M.R., Lowry, G.V., Alvarez, P., Dionysiou, D., and Biswas, P. (2006). Assessing the risks of manufactured nanoparticles. Environ. Sci. Technol., 40, 4336-4345. Wigginton, N.S., Haus, K.L., and Hochella, M.F. (2007). Aquatic environmental nanoparticles. J. Environ. Monitor., 9, 1306-1316. Wilkinson, K.J., and Lead, J.R. (2007). Environmental Colloids and Particles: Behaviour, Separation and Characterisation. John Wiley and Sons, Chichester, UK. Wittmaack, K. (2007). In search of the most relevant parameter for quantifying lung inflammatory response to nanoparticle exposure: particle number, surface area, or what? Environ. Health Perspect., 115, 187–194. Wolfrum, E.J., Huang, J., Blake, D.M., Maness, P.C., Huang, Z., Fiest, J., and Jacoby, W.A. (2002). Photocatalytic oxidation of bacteria, bacterial and fungal spores, and model biofilm components to carbon dioxide on titanium dioxide-coated surfaces. Environ. Sci. Technol., 36, 3412-3419. WWIC (2003). Nanotechnology and regulation. A case study using the Toxic Substance Control Act (TSCA). Woodrow Wilson International Centre for Scholars Foresight and Governance Report. Xu, X-H.N., Brownlow, W.J., Kyriacou, S.V., Wan, Q., and Viola, J.J. (2004). Real-time probing of membrane transport in living microbial cells using single nanoparticle optics and living cell imaging. Biochem., 43, 10400-10413. Xu, Z.P., and Lu, G.Q. (2006a). Layered double hydroxide nanomaterials as potential cellular drug delivery agents. Pure Appl. Chem., 78, 1771-1779. Xu, Z.P., Stevenson, G., Lu, C. Q., and Lu, G.Q. (2006b). Dispersion and size control of layered double hydroxide nanoparticles in aqueous solutions. J. Phys. Chem. B, 110, 16923-16929. Yamamoto, O. (2001). Influence of particle size on the antibacterial activity of zinc oxide. Intern. J. Inorg. Mater., 3, 643-646. Yang, L., and Watts, D.J. (2005). Particle surface characteristics may play an important role in phytotoxicity of alumina nanoparticles. Toxicol. Lett., 158, 122-132.

Zheng, L., Hong, F., Lu, S., and Liu, C. (2005). Effect of nano-TiO2 on strength of naturally aged seeds and growth of spinach. Biol. Trace Element Res., 104 , 83-91.

Zhu, S., Oberdörster, E., and Haasch, M.L. (2006). Toxicity of an engineered nanoparticle (fullerene, C60) in two aquatic species, Daphnia and fathead minnow. Mar. Environ. Res., 62, S5–S9.

Fate of Manufactured Nanomaterials in the Australian Environment 72

Zhu, X., Zhu, L., Li, Y., Duan, Z., Chen, W., and Alvarez, P.J.J. (2007). Developmental toxicity in zebrafish (Danio rerio) embryos after exposure to manufactured nanomaterials: buckminsterfullerene aggregates (nC60) and fullerol. Environ. Toxicol. Chem., 26, 976-979.

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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 membrane-bound 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 physico-chemical 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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