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Fate of Manufactured Nanomaterials in theAustralian Environment

G.E. Batley and M.J. McLaughlinCSIRO Niche Manufacturing Flagship Report

March 2010

Prepared for the Department of the Environment, Water, Heritage andthe Arts

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Fate of Manufactured Nanomaterials in the Australian Environmentii

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Graeme BatleyCentre for Environmental Contaminants ResearchCSIRO Land and WaterPrivate Mailbag 7, Bangor NSW 2234Phone 02 9710 6830Fax 02 9719 6837Email [email protected]

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Fate of Manufactured Nanomaterials in the Australian Environment iii

EXECUTIVE SUMMARY

With growing production and use of manufactured nanoparticles in a large range of consumerproducts, regulatory agencies worldwide are addressing the risk that these substances maypose to both the environment and human health. An assessment with respect to ecosystemhealth requires an ecological risk assessment that must take into account current knowledgeabout nanomaterial uses, environmental concentrations, fate, and effects, to determine bothpredicted environmental concentrations (PECs) and predicted no-effect concentrations(PNECs).

This report reviews the available literature on the fate of manufactured nanomaterials in theaquatic 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 areaggregation and dissolution, driven by size and surface properties of the materials. Theseprocesses can be mediated by interactions with dissolved organic matter and other naturalcolloids. Biological degradation processes and abiotic degradation via hydrolysis andphotolysis do not appear to be significant in waters, although oxidation/reduction reactionscan be significant for some metals.

Similar processes are operative in terrestrial systems, but mobility is much reduced comparedto aquatic environments. Interactions of nanoparticles with soil minerals and organic matterhave not been evaluated, but are likely to be a function of particle size, shape and surfaceproperties (specific surface area and surface charge). Small hydrophilic nanoparticles (<20nm) with net negative surface charges are likely to be mobile, while large hydrophilicpositively-charged particles will be sorbed by soil. Strongly hydrophobic nanoparticles arelikely to be strongly retained by soil organic matter.

Many parallels can be seen in the behaviour of natural colloids. In considering the behaviourof manufactured nanomaterials, it is important that studies be carried out in natural waters asthe often orders of magnitude higher concentration of natural colloids can have a significantimpact. Aggregation results in growth of nanoparticles, often to sizes in excess of thenanoparticle size definition of <100 nm, ultimately leading to sedimentation. This growth canbe prevented by the presence of surfactants and other surface coatings, or through the presenceof natural humic materials. Fibrillar colloids enhance precipitation. Any toxicity studies willneed to separately address particular nanomaterial formulations.

There is evidence to suggest that the impact of manufactured nanoparticles on aquaticorganisms differs compared to their macroparticle equivalents. In some instances suchassessments 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, physicaldamage 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 Environmentiv

across the gill and other external surface epithelia for higher organisms. Bioaccumulation viathe food chain is also possible. The extensive literature on bacterial toxicity was consideredinappropriate for defining no effects concentrations of nanomaterials in waters, however, theeffects, both positive and negative, of some nanomaterials on bacteria present in sewagetreatment plants and on soil microorganisms have been discussed.

There are limited data on toxicity of nanoparticles to algae, invertebrates and fish. In the caseof n-C60 fullerenes, toxicity was highly dependent on the method of preparation, with theparticles dispersed by evaporation of tetrahydrofuran (THF) extracts being more toxic thanthose dispersed by sonication, due supposedly to secondary effects of THF. Insufficient datawere available to derive high reliability environmental guidelines, but for freshwaters, lowreliability guidelines were derived for n-C60 and TiO2 nanoparticles. The calculated PNECvalues are only marginally above the concentrations estimated to be released to theenvironment in calculations based on nanomaterial usage in the UK.

There is much less information on the behaviour and toxicity of nanoparticles in terrestrialsystems, due to difficulties in assessing dose against a background of natural nanoparticles inthe soil matrix. Heterogeneity and incorporation of nanoparticles into soil is also an issue forecotoxicological testing. There are a few reports of adverse effects of some nanoparticles toterrestrial species cultured in vitro, but to date there is no strong evidence that nanomaterialshave significant adverse effects on terrestrial species in soil exposures. Further studies areneeded with a wide range of terrestrial species, and a wide range of nanoparticulate materialsin a range of soil environments, to determine if the preliminary data are sound.

There are numerous examples to demonstrate that nanomaterials can be bioaccumulated byorganisms. The extent to which this uptake exerts toxicity is less certain.

Current international activities in relation to nanoparticle risk assessment are discussed. Insummary, most countries see the need for more data gathering and research to improve therisk assessment of these materials. This review indicates that the same is possibly true forAustralia, but the way ahead is reasonably clear. The basic recommendations for futureresearch are:

1. There is a need for measurements in natural water, sediment and soil samples of thestability, and short- and long-term fate of the various likely formulations that mightreach these compartments of the environment. As well, techniques are needed todistinguish natural from manufactured nanoparticles. These measurements shouldfocus 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 asrequired to derive PNECs using species sensitivity distributions. It is critical thatappropriate verification of particle and solute dose be undertaken in all ecotoxicitytesting, 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

Fate of Manufactured Nanomaterials in the Australian Environment v

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

4. Comparisons of toxicity testing in natural vs. synthetic soil and water samplesdemonstrating the effects of natural colloids.

Understanding the fate of nanoparticles in the Australian environment will assist risk assessments byguiding the toxicity testing of nanomaterial formulations under real environmental conditions,yielding realistic PNECs. This should be coupled with the development of appropriatemeasurement techniques that can quantify both concentrations and particle sizes withappropriate quality assurance and quality control. As well as size and composition, it isevident that surface properties of nanoparticles will be fundamental in determining fate andtoxicity in the environment and these properties will need to be considered in any hazardranking.

A check list has been provided to incorporate fate considerations in assessing bothenvironmental exposure and effects of manufactured nanomaterials.

Fate of Manufactured Nanomaterials in the Australian Environmentvi

Contents

EXECUTIVE SUMMARY ............................................................................................. iii

1. INTRODUCTION .................................................................................................1

2. CLASSES OF NANOMATERIALS ......................................................................2

3. NANOPARTICLE USAGE IN AUSTRALIA..........................................................7

4. CHARACTERISTICS OF ENVIRONMENTAL NANOPARTICLES ......................94.1 Manufactured Nanomaterials ....................................................................................9

4.2 Natural Nanoparticles................................................................................................9

5. ENVIRONMENTAL SOURCES OF MANUFACTURED NANOPARTICLES .....12

6. FATE OF NANOMATERIALS IN AQUATIC SYSTEMS ....................................146.1 Key Pathways..........................................................................................................14

6.2 Behaviour of Manufactured Nanoparticles ..............................................................156.2.1 Aggregation ........................................................................................................ 156.2.2 Nanoparticle Solubility ........................................................................................ 176.2.3 Role of Nanomaterial Formulations and Impurities............................................. 196.2.4 Fate in Natural Water Systems........................................................................... 206.2.5 Nanoparticles as Vectors for Contaminant Transport ......................................... 21

6.3 Fate of Manufactured Nanomaterials in Terrestrial Systems..................................226.3.1 Key Pathways ..................................................................................................... 226.3.2 Behaviour of Natural Colloids in Soils................................................................. 236.3.3 Behaviour of Manufactured Nanoparticles in Soils ............................................. 24

7. ECOLOGICAL RISK ASSESSMENT OF MANUFACTURED NANOPARTICLES.......................................................................................................................... 257.1 Polymeric Nanoparticles as a Separate Class ........................................................26

8. EXPOSURE ASSESSMENT .............................................................................268.1 What to Measure.....................................................................................................26

8.2 Methods for Measurement of Nanoparticles ...........................................................278.2.1 Relevance of OECD Test Guidelines.................................................................. 28

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

9. ECOTOXICOLOGY OF NANOPARTICLES......................................................339.1 Ecotoxicity and Nanoparticle Dose Metrics .............................................................33

9.2 Toxicity to Aquatic Biota ..........................................................................................359.2.1 Mechanisms of Biological Uptake and Toxicity................................................... 359.2.2 Ecotoxicity to Individual Species......................................................................... 369.2.3 Developing Appropriate Guidelines for Nanomaterials in Waters ...................... 439.2.4 Bioaccumulation ................................................................................................. 449.2.5 Ecological Impacts.............................................................................................. 45

Fate of Manufactured Nanomaterials in the Environment vii

9.3 Sediment Toxicity.................................................................................................... 46

9.4 Toxicity to Terrestrial Biota...................................................................................... 469.4.1 Ecotoxicity to Individual Species .........................................................................469.4.2 Development of Guidelines for Nanomaterials in Soils .......................................49

10. INTERNATIONAL PROGRESS ON NANOMATERIAL RISK ASSESSMENT...4910.1 International Approaches ........................................................................................ 49

10.1.1 USA .....................................................................................................................4910.1.2 United Kingdom...................................................................................................5110.1.3 Other International Activities................................................................................52

10.2 Australian Activities ................................................................................................. 53

11. DEVELOPMENT OF TECHNICAL GUIDELINES FOR NANOMATERIALASSESSMENT.................................................................................................. 5411.1 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 Environmentviii

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 ofsurfactants 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 Müllenand 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 intowhich 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 aZnO 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 fromBjorn et al., 2006)

Figure 9. Key processes in soil relating to transformation and potential risk from manufacturednanoparticulate particles

Figure 10. Framework for deriving mass flow data for silver flows from biocidal plastics and textiles (fromBlaser 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 Boxallet al., 2007)

Table 6. Predicted environmental concentrations (PEC) of nano-Ag, nano-TiO2 and CNTs in air, waterand 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)

Fate of Manufactured Nanomaterials in the Australian Environment ix

Table 8. Approach to toxicity testing of nanomaterials in waters

Table 9. Summary of toxicity testing results for manufactured nanomaterials (expanded from Apte et al.,2008)

Table 10. Data for estimation of guideline concentrations for n-C60 in freshwater

Table 11. Published evidence of nanoparticle uptake by aquatic organisms (from Apte et al., 2008)

Table 12. Toxic effects of nanomaterials on soil organisms (from Klaine et al., 2008)

Fate of Manufactured Nanomaterials in the Australian Environment 1

1. INTRODUCTION

The last decade has seen an amazing growth in nanoscale science and technology, to theextent that nanomaterials are now a component of a wide range of manufacturedproducts, from sunscreens to sensors. Given that the production volumes of some ofthese materials is already exceeding thousands of tonnes, there is growing public andregulatory concern for the potential adverse effects that release of these materials to theenvironment may have on both human and ecosystem health. Nanoparticles are alreadypresent 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 toaccount for more than 36–44% of the total number of particles in some urban airsamples, 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 atmosphericenvironment (usually referred to as fine and ultrafine particles) have been well studiedand there are clear concerns for the finer particles that can reach the deeper recesses ofthe lungs. For terrestrial and aquatic environments, there has been extensive research onnatural colloids (Buffle and Leppard, 1995a), however, there have been few studies ofanthropogenic particles.

Manufactured nanomaterials can be defined as those that are deliberately producedrather than materials that are by-products of other activities not targeted at nanomaterialproduction. Nanomaterials are commonly based on nanoparticles, for which theaccepted definition is particles that have at least one dimension less than 100 nm, butthe term is also used to refer to materials such as surfaces with nanometre-sized featuresthat are not particulate in nature, or substances with nanometre size voids. Small sizegives materials properties that differ from those of bulk or macroscopic materials. Inparticular, optical, electrical and magnetic properties can differ in ways that are subjectto the laws of quantum rather than classical physics. Nanoparticles have a large surfaceto volume (and mass) ratio, and potentially greater reactivity and mobility. Surfaceareas 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 distinctivenano properties, although manufacturers are devising coatings that can stabilisenanoparticles. Smaller size carries with it the potential to be more bioavailable, able topenetrate biological membranes or to enter cells by endocytosis (engulfing by the cellwall).

This report reviews the current state of knowledge with respect to nanoparticle fate andeffects in the environment, with a particular focus on aquatic and terrestrial systems, toprovide a foundation for the risk assessment of manufactured nanoparticles in Australia.

Fate of Manufactured Nanomaterials in the Australian Environment2

2. CLASSES OF NANOMATERIALS

Manufactured nanomaterials currently fall into one of at least seven different classes, asshown 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 thatcapitalise on their enhanced properties. A case in point is zinc oxide that has been usedfor 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, plasticsand packaging

Titanium dioxide Cosmetics

Cerium dioxide Automobile catalyst

Mixed oxides Cosmetics

Carbonproducts

Fullerenes

Single-walled andmulti-walledcarbon nanotubes

Plastics, catalysts, battery and fuel cell electrodes,super-capacitors, water purification systems,cosmetics, orthopedic implants, conductive coatings,adhesives and composites, sensors, and components inelectronics, aircraft, aerospace and automotiveindustries

Amorphouscarbon

Inks, photocopier toner, automobile tyres

Metals Silver Bactericide in wound dressings, socks and othertextiles, air filters, toothpaste, baby products, vacuumcleaners, and washing machines

Iron Remediation of groundwaters, sediments, soils

Gold Electronics in flexible conducting inks or films, and ascatalysts

Bimetallicnanoparticles Fe-Pd, Fe-Ni, Fe-Ag

Remediation of organics in waters; usually supportednanoparticles

Quantum dotsandsemiconductors

CdTe, CdSe/ZnS,CdSe, PbSe andInP

Medical applications, photovoltaics, security inks, andphotonics and telecommunications

Nanoclays Hectorites,layered doublehydroxides

Cosmetics, toothpaste, antacids, paint additives,catalyst supports, flame retardants, drug deliveryagents

Dendrimers Coloured glasses, chemical sensors, and modified


electrodes; in medicine as DNA transfecting agents,therapeutic agents for prion diseases, formation ofhydrogels, drug delivery, DNA chips, and ex vivoamplification of human blood cells

Emulsions Acrylic latex andotherformulations

Paints and surface coatings; sunscreens and similarcream formulations; in medicine for drug delivery;pesticide formulations

In a nanoparticulate form, ZnO is transparent to the eye, but retains much of its ability toabsorb UV radiation, albeit over a narrower spectrum. In Australia in 2005, of the 1200sunscreens authorised by the Therapeutic Goods Administration (TGA), 228 containedzinc oxide, 363 contained titanium dioxide and 73 contained both (TGA, 2006).

Nano-zinc oxide coatings on clear glass beer bottles prevent UV-degradation of thecontents, while making them appealingly visible to the consumer. Other metal oxides incommon use include titanium and cerium dioxides, while mixed-metal compounds suchas indium-tin oxide (ITO) are currently used in polishing agents for semiconductorwafers, sunscreen formulas and scratch-resistant coatings for glass (Arabe, 2003).

Carbon-based nanoparticles comprise the second class (Figure 1). This includesfullerenes, carbon nanotubes (CNTs) and amorphous carbon nanoparticles. The firstfullerene was discovered in 1985, a sixty carbon atom hollow sphere known as thebuckyball was produced by evaporating graphite (Kroto and Walton, 2007). It wasrecently revealed that naturally-produced fullerenes have been around for over a billionyears, found in parts per million concentrations in ancient rock formations and believedto 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 canbe synthesised under controlled conditions to a particular diameter and size, either fromgraphite using an arc discharge or laser ablation, or from a carbon-containing gas using


chemical vapour deposition. The multiwalled-products (MWCNTs) are concentriccylinders up to 10 nm in length and 5–40 nm in diameter. It was later shown that it waspossible 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 ofsteel (Lekas, 2005).

In aqueous systems, carbon nanoparticles aggregate, due to their inherenthydrophobicity. This limits their use in aqueous and biomedical applications. Muchresearch has been done to surface modify these nanoparticles to stabilize aqueoussuspensions. Covalent modification, such as the attachment of polyethylene glycol toSWCNTs (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 haveimplications for their use in certain applications as well as repercussions for their fateand 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 ofsurfactants 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 factorieswith 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 inplastics, catalysts, battery and fuel cell electrodes, super-capacitors, water purificationsystems, orthopedic implants, cosmetics, conductive coatings, adhesives andcomposites, sensors, and components in electronics, aircraft, aerospace and automotiveindustries. Increased production results in an increased potential for release to theenvironment, either deliberately in discharges or accidentally in spillages, and a greaterpossibility of adverse environmental effects. Increased manufacturing volumes alsoincrease the absolute quantities discharged to the environment as a result of use inproducts, and significantly, the quantities that must be disposed of.

www.f-

www.fullereneinternational.com


The third class comprises nanoparticulate zerovalent metals such as silver, gold andiron. Nanoparticulate zerovalent iron has been used for some time for the remediationof waters, particularly groundwaters, as well as sediments and soils (Tratnyek andJohnson, 2006). It has been used to remove nitrates via reduction, and has most recentlyfound use in detoxifying organochlorine pesticides and polychlorinated biphenyls(Zhang et al. 2003). Mobile iron nanoparticles are effective in treatment of dissolvednon-aqueous phase liquids (DNAPL) (Tratnyek and Johnson, 2006).

Bimetallic nanoparticles such as Fe-Pd, Fe-Ni, Fe-Ag, Pd-Ru, etc., have found extensiveuse 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 consumerproducts, as indicated in the inventory developed by the Woodrow Wilson InternationalCentre for Scholars Project on Emerging Nanotechnologies (PEN, 2007a). Applicationsare largely based on its bactericidal activity, and include wound dressings, socks andother textiles, air filters, toothpaste, baby products, vacuum cleaners, and washingmachines. In some cases, the active ingredient is metallic nanoparticulate silver, inothers, ionic silver (Ag+) is electrochemically generated. Ionic silver is not really ananoparticle, but is highly particle reactive, so in natural waters is readily adsorbed byboth macroparticles and by colloidal particles such as iron oxyhydroxides or naturalorganic 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 hasentirely different properties as a non-ionic nanoparticle. In both cases, however, theinstability of the monovalent cation and the non-ionic nanoparticle result in extremelyshort half-lives of the desired form. This has resulted in research to stabilise silvernanoparticles to make them useful in biological and other aqueous applications (Doty etal., 2005). This has created ambiguity in how investigators describe test systems andmanufacturers describe products. For example, it is common for manufacturers todescribe colloidal silver as ‘nanosilver’, rather than metallic silver powder that iscommercially available as nanoparticles.

Colloidal elemental gold has been used for many years, especially in medicalapplications as a vector in tumour therapy. Its size varies from 20-160 nm and thespectral properties change with the classical colour variation from ruby red throughpurple to pale blue as size increases (Turkevich et al., 1954). Newer applications ofnanoparticulate 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 afourth class of nanomaterials. Typical materials include CdTe, CdSe/ZnS, CdSe, PbSeand InP with size ranges from 10 to 50 nm. They usually consist of a semiconductorcore surrounded by a shell, e.g. silica (Sass, 2007). Newer formulations are comingonto 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 aparticle of light. Changing the size of a QD-based LED makes it emit a differentwavelength of light – producing red, orange, yellow, or green light. The devices areuseful in that they only need about 3 to 4 volts to operate and can run for over 300 hourswithout losing any brightness. Their surface is usually functionalised by coatings toensure solubility in water.

Synthetic clays represent a large class of nanomaterials with over 9000 tonnes ofnanoclays being produced in 2007 (Electronics.ca, 2007). Both manufactured andnatural clays are starting materials in nanocomposites for use in polymernanocomposites, in packaging, paints and cosmetics. They typically range in size from10 nm to 100 nm. Both negatively charged and positively charged platelets can beobtained. The former include montmorillinite and hectorite clays, while syntheticlayered double hydroxides (LDHs) of magnesium and aluminium have exchangeableinterlayer anions (Choy et al., 2000; Xu et al., 2006a.b). The anion exchange propertiesof these materials allow binding to negatively charged biomolecules between thehydroxide layers, with hybridisation effectively neutralising the charge and facilitatingpenetration into cells, hence their potential as drug delivery agents.

Dendrimers are monodisperse multifunctional polymers that have repeatedly branchedcomponents that form a fifth class of nanomaterials. They are typically spheroid orglobular nanostructures designed to carry molecules encapsulated in their interior voidspaces or attached to their surface (Figure 3). They range in size from around 5 nm forthe 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 tothe nanoscale (e.g. see Frechet and Tomalia, 2002; Dendritic Technologies Inc,www.dnanotech.com). These macromolecules can be used for many useful applicationsin different fields from biology, material sciences, surface modification, toenantioselective catalysis.

Figure 3. Typical structure of a dendrimer (first-generation polyphenylene dendrimer reported by Müllenand coworkers in Chem.-Eur. J., 2002, 3858-3864)

www.dnanotech.com


Among the most outstanding applications of dendrimers are the formation of nanotubes,micro and macrocapsules, nanolatex, coloured glasses, chemical sensors, and modifiedelectrodes. 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 ofnanoparticles, more recently referred to as soft nanoparticles. Emulsions are bydefinition dispersions of one immiscible liquid in another and while we are accustomedto thinking of particles as solid phases, the term really refers to ‘small amounts’ socould include emulsions. A nanoemulsion has been defined as a type of emulsion inwhich the sizes of the particles in the dispersed phase are less than 1000 nm. Thisincludes particles much larger than the accepted nanoparticle size of < 100 nm, andtypically 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 fordrug delivery. It is worth noting that the term microemulsion is also used to describeoil/water emulsions in the nanoemulsion size range and below. The distinction is a fineone, with microemulsions being thermodynamically stable while nanoemulsions arekinetically stable (Lawrence and Waresnoicharoen, 2006).

From the list of manufactured nanoparticles and their reported uses, apart from the useof iron and related bimetallic nanoparticles for water and soil remediation, it appearsthat there are few confirmed uses of nanoparticles as agricultural or veterinarychemicals. So saying, there is potential for use in veterinary medicine for drug deliveryuses and other applications common to human medical uses. One reference highlightedthe use of nanoemulsions for crop applications(www.nanowerk.com/spotlight/spotid=5305.php).

A distinction has been made by some authors between nanosized particles andnanosized molecules. The latter include fullerenes and dendrimers. If a moleculecontains 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 functionalsignificance of these separate definitions is not immediately obvious. There will bedifferences in fate and toxicity just as there are between different types of non-molecularnanoparticles.

3. NANOPARTICLE USAGE IN AUSTRALIAA voluntary call for information on nanoparticle usage in Australia was recently issuedby the National Industrial Notification and Assessment Scheme (NICNAS). In theabsence of publicly available information, the call was targeted at manufacturers andimporters of nanomaterials or products containing nanomaterials for industrial(including domestic and cosmetic) use during 2005 and 2006. It was designed as a firststep in understanding the potential for exposure. In addition to issuing an open call inthe Chemical Gazette of February 2006, companies known to be involved withnanomaterials were directly contacted by NICNAS. The results of the survey were

www.nanowerk.com/spotlight/spotid


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

Table 2. Usage of nanomaterials in the commercial sector in Australia

Chemical Name Application Total usage (tonne/y)

Acrylic latex Surface coatings 10,000-50,000

Aluminium oxide Printing 0.05-0.1

Aluminosilicates Water treatment 10-50

Carbon black pigment Surface coatings 10-50

Cerium oxide Catalysts 1-5

Iron oxide Surface coatings

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 ofacrylic latex in nanoemulsions, was the fact that CNTs, fullerenes and silver were notimported or manufactured (as chemicals or in products) at that time, given the highproduction volumes projected for 2008-2009 internationally. A second survey iscurrently 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 ofnanotechnology-based consumer products (PEN, 2007a) lists 45 carbon-based products

www.nicnas.com.au


as of February 22, 2008. This is the second most common material after silver (143references), 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 weknow it is a component of many consumer products, or to nanoclays. Further localresearch is needed to confirm actual usage of CNTs, silver and nanoclays.

4. CHARACTERISTICS OF ENVIRONMENTALNANOPARTICLES

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 theirchemical composition. Measurement of these properties is not a trivial exercise as willbe discussed later. The challenge is to determine these properties when thenanomaterials reach the particular environmental compartments, (atmospheric, aquatic,terrestrial). Where the measurement technique requires the nanoparticles to beseparated, e.g. electron microscopy, the possibility exists that this will perturb thenatural properties from their form in the environment.

4.1 Manufactured Nanomaterials

The properties of manufactured nanomaterials, as produced, will vary greatly oncereleased to the environment as interactions occur with other chemicals, or astransformations such as aggregation and dissolution take place. Such processes candramatically affect subsequent biological interactions. Because of this, the concept ofintrinsic toxicity of manufactured nanoparticles is not a useful one, and needs to belinked with measurements in field or simulated field media. There is a parallel here inthe study of metal speciation in aquatic and terrestrial systems, where the guidelineframework uses a conservative, total dissolved metals as a first cut before a detailedmeasurement of a bioavailable fraction. Here the conservative assumption might be tofirst base assessments of biological impact on the smallest and potentially morebioavailable particles in the absence of later more appropriate measurements of actualsize.

The situation becomes more complex because the formulations of manufacturednanomaterials can often include other additives that can alter the physical behaviour ofnanoparticles in some media, as will be discussed later. Such a concept is notunfamiliar, 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 thereare a range of natural nanoparticles. In air, there are dusts as well as aerosolscomprising 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 byagglomeration of much smaller particles. Natural clays can be a significant


nanoparticulate component of some soils, as can iron and manganese oxides and otherhigh molecular weight mineral phases as well as dissolved organic matter in soil porewaters. In natural waters, as well as in soil pore waters, colloidal particles compriseclays, iron and manganese oxides and organic matter. We can learn a lot about theexpected behaviour of manufactured nanoparticles from what is already known aboutnatural 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 andaluminium oxides (Wilkinson and Lead, 2007). Natural colloids fall in the size rangefrom 1–1000 nm, depending on their degree of aggregation (Buffle and Leppard, 1995a;Lead and Wilkinson, 2007). Fulvics, humics, and proteins are typically smaller thantens of nm while polysaccharides and metal oxides are larger, although iron oxidecolloids cover the full size range. Typically these are not present as discrete particles orcompounds, but are heterogeneous mixtures of organic and inorganic species (Lead andWilkinson, 2007). Microbial activity in natural waters is a continuous source ofmacromolecular material (e.g. polysaccharides) (Buffle and Leppard, 1995b)

Over the wide range of colloid particle sizes, the largest particles have the greatestpercentage mass, but the smaller particles have the greatest number and percentage oftotal surface area. Buffle and Leppard (1995a) showed that irrespective of the aquaticsystem of interest, the size distribution based on particle number (N) follows Pareto'sLaw (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 (particlenumber/particle diameter) and log (diameter) means that there are orders of magnitudemore smaller particles than large ones in a water system.

The aggregation of colloids is dependent upon particle size, density, surface charge andchemical properties (Buffle and Leppard, 1995a; Handy et al., 2008). Aggregationoccurs as a result of particle-particle collisions, involving natural Brownian motion,different shear velocities in flowing systems and different settling behaviour of differentsized particles.

It has been shown both practically and theoretically, that for a mixture of colloids inwhich each size fraction has the same volume, the smallest colloids (<100 nm)disappear first by aggregation, and the largest by sedimentation, leaving a distribution ofsizes over the range 100 nm-1 µm. (Buffle and Leppard, 1995b).

The interactions between colloids will be governed by their charge and the nature oftheir bonding (covalent vs electrostatic). The surface charge of clays at the pH ofnatural waters is typically negative over a range of natural pH values. So too is thecharge on most natural organic matter due to ionisable functional groups (e.g. hydroxyland carboxylic acid). Iron and aluminium oxyhydroxides have a positive charge belowthe pH values at which the surface charge is zero (pH 8-9), however, binding withnatural organic matter typically results in aluminium and iron colloids having a netnegative charge in natural waters (Kretzschemar and Schafer, 2005).


It is not easy to measure surface charge, but it is implied by measurements of the zetapotential (the potential between the colloid particle surface and solution). As a measureof the stability of colloidal particles, the zeta potential range between +30 mV and -30mV is characterised by instability with aqueous dispersions being stable on either sideof that range.

Particles with near neutral charges aggregate rapidly. In natural systems, suchinteractions of organic macromolecules and colloidal particles lead to the formation ofloose aggregates or flocs whose structure will be dependent on the relativeconcentrations of each in the mix and of the density, shape of the particles and theflexibility of the macromolecules. Their stability depends on their relative charges andthe nature of the bonding. The aggregates may be stabilised at small sizes that will notsediment. Larger aggregates form more slowly. It is difficult to predict the behaviourof such mixtures in terms of rates of reaction and stability, however, it appears that inlow ionic strength solutions, appreciable stability is generally achieved in the size range100 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 rateof settling will be controlled by both hydrodynamics and particle size. Deeper, well-mixed waters have reduced settling and larger colloidal aggregates. Once the aggregatesbecome sufficiently large (>1 µm) they exceed the buoyant mass of colloids and thenewly 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 byorganic macromolecules (Gustafsson and Gschwend, 1997). In estuarine waters,increasing ionic strength increases screening of the particle charge, resulting inincreased aggregation and coagulation of colloidal particles (Buffle and Leppard,1995a). For example, the aggregation and precipitation of colloidal iron at salinitiesabove 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 andMorgan, 1990).

A schematic diagram of aggregate formation involving natural colloids is shown inFigure 4 (from Buffle et al., 1998). This does not consider any living components suchas bacteria and viruses which would add a further layer of complexity.

Natural colloids are frequently in high concentrations in soil pore waters and in naturalwater systems, as high as mg/L, so interactions of these particles with manufacturednanoparticles will be an important fate pathway to consider, and one that is overlookedin laboratory investigations in synthetic media. The basic behaviour of natural colloidsand macromolecules in soils has been known for decades (Cameron, 1915), and isgoverned by the same processes as those in natural waters. High ionic strength (saltcontent) in soils will promote flocculation of particles, as will soil pore watersdominated by calcium and low in sodium (Rengasamy and Olsson, 1991). ManyAustralian soils are sodic (sodium rich) (Naidu and Rengasamy, 1993), conditionswhich 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 adversesoil 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 similarlyon manufactured nanoparticles, although this needs confirmation.

Figure 4. Major types of aggregates formed in the three-colloidal component system: fulvic compounds(or aggregated refractory organic material), small points; inorganic colloids, circles; rigid biopolymers,lines. Both fulvics and polysaccharides can also form gels, which are represented here as gray areas intowhich inorganic colloids can be embedded. (From Buffle et al., 1998)

5. ENVIRONMENTAL SOURCES OF MANUFACTUREDNANOPARTICLES

Unlike many anthropogenically derived nanoparticles, it is reasonable to assume thatthere will controls on the release of manufactured nanoparticles that minimise theirrelease to the environment. The obvious sources that require management are release tothe atmosphere and release via aqueous discharges. These and other potential inputsources are illustrated schematically in Figure 5.

Atmospheric nanoparticles are both a potential risk to the environment and anoccupational health and safety concern for workers engaged in nanomaterialmanufacture. Sources include motor vehicle exhausts and stack emissions from a rangeof sources. Eliminating exposure to workplace respirable nanoparticles will be apriority and is easily addressed through both the use of filters and appropriate protectiveclothing and respiratory protection. Motor vehicle and stack emissions are moreproblematic. Dealing with fine particle emissions has been an issue for the powerindustry for many years, and it is fair to say, has not been adequately eliminated.


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

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 liquidwastes that may contain nanoparticles. These days, discharges of aqueous wastes arelicensed, although there is unlikely to yet be advice on nanoparticle concentrations.Similarly uncontrolled disposal of solid chemical wastes is generally not permitted, butguidance on nanomaterials is most likely absent, so this remains a potential source.

Water treatment plants may have the capability to treat and remove nanomaterials wheredischarges are to the sewage system, but as yet there is no information on the ability ofwater treatment plants to deal with nanoparticulate contaminants. In particular, anionicand uncharged nanomaterials could pass through into sewage effluents and not beretained in sewage biosolids. Several recent studies have indicated a potential fornanomaterials to interact with bacteria in sewage treatment plants. Choi et al. (2008)showed that silver nanoparticles were toxic to nitrifying bacteria and that this couldimply detrimental effects on the microorganisms in wastewater treatment. Titaniumdioxide 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 etal. (2008) found that SWCNTs caused the protozoan Tetrahymena thermophilia, presentin sewage treatment plants to release excess exudates, which contribute to flocformation, so they could be used to improve the efficiency of ciliates in wastewatertreatment although effective measures to control and monitor SWCNT release would benecessary. By contrast, Nyberg et al. (2008) recently indicated little toxicity offullerenes in sewage treatment sludge to methanogenic bacteria.

Atmospheric deposition

Surface runoff

Effluent discharge

Groundwater dischargeRoad runoff

Accidental spillage

Soil application


There are instances where nanomaterials are added to aquatic or terrestrial systems forremediation purposes, e.g. zerovalent iron addition to soils or sediments, and their fateand impacts will be separately discussed.

The remaining sources are accidental release and release as a consequence of productusage. The accidental release amounts to spillage of containers, drums, etc., where solidnanomaterials can end up on land or in water systems. The issue of product usagerequires consideration for each nanomaterial category and for formulations within eachcategory and this will be discussed in more detail below.

The US Department of Energy has recently published an approach to nanomaterialenvironmental safety and health that discusses in some detail the requirements fornanomaterial transportation and for the management of nanomaterial-bearing wastestreams and nanomaterial spills that minimise the likelihood of releases ofnanomaterials to the environment (USDOE, 2007).

In addressing the risks posed by manufactured nanomaterials, a relevant question iswhich nanoparticles have the highest potential for release. Intuitively, these are likely tobe those being produced in the greatest amounts, however, if these productions are froma large number of widely-dispersed small scale activities, perhaps the risk is less thanfrom larger facilities with high volume throughputs. Silver fits into the small butdispersed source category. In particular, the in situ generation of silver nanoparticles inwashing machines will be a highly dispersed source, that may end up in wastewatertreatment plants, and from there may reach the environment although may well berecovered in the flocculation stages of such plants.

The particular formulation of the nanomaterials is also important for assessing potentialfor diffuse releases into the environment. Where nanoparticulates are incorporated intostable solid-phases, e.g. ZnO nanoparticles in coatings on glass for UV protection, thenthe potential for release of the dispersed nanoparticles is low. Where the nanoparticle isused in a dispersed form (e.g. zerovalent iron for groundwater remediation), then thepotential 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 theaquatic environment are summarised in Figure 6. These comprise aggregation andsubsequent sedimentation, dissolution, adsorption to particulates and other solidsurfaces, binding to natural dissolved organic matter, and stabilisation via surfactants.Other processes include biological degradation (aerobic and anaerobic), and abioticdegradation (including hydrolysis and photolysis). Oxidation and reduction may also beof concern in some environments for specific materials. Concentration in the surfacemicrolayer of water bodies is a possibility, but unlikely to be a major pathway. Theultimate fate is likely to involve accumulation and burial in bottom sediments.


In general, the fate of manufactured nanomaterials in aquatic systems has not been thatwell studied, however, what information is available, coupled with the extensiveliterature on natural colloids in aquatic systems can provide a useful basis for predictionof nanomaterial fate. The interactions of nanomaterials with natural colloids will play acritical 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 theenvironmental impacts of manufactured nanomaterials in waters are aggregation anddissolution.

6.2.1 Aggregation

As shown for natural nanoparticulate colloids (in Section 4.2), the behaviour ofnanoparticles in aqueous systems mimics colloid behaviour. There is a naturalpropensity for nanoparticles to grow in size in aqueous solution. Particles thataccording to manufacturers’ specifications are nanosized, when suspended in water atneutral pH, are frequently aggregated (e.g. Figure 7), and the size of these aggregates isfrequently 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 byelectrolyte ions, e.g. in seawater, overcomes the electrostatic forces and allowsaggregation, as for natural colloids. Steric stabilisation of nanoparticles against

Binding to NOM

and other colloids

Aggregation

Sedimentation

Binding to NOM

and other colloids

Mn+

Dissolution

Mn+

Mn+

Surfactant-stabilisednanoparticles

Binding to suspendedparticles/biota

Biological degradation,photolysis, hydrolysis


aggregation can occur through surface modification by surfactants or bulky polymericadditives.

Interactions of nanomaterials with natural colloids (organic macromolecules, inorganiccolloids or heterogeneous aggregates) will also occur in the same manner as discussedin Section 4.2 (Saleh et al. 2008).

The rates at which manufactured nanoparticles aggregate is particularly important, sincethe slower the aggregation the greater the potential for interaction with biota.Unfortunately this has been poorly studied, although there are data available for naturalcolloidal nanomaterials.

The terms aggregate and agglomerate have distinct meanings in particle science, but arefrequently confused. As discussed by Nichols et al. (2002), agglomerates are generallyconsidered to be an assemblage of particles that are rigidly bound by fusion sintering orgrowth, while aggregates are loosely bound particles that are readily dispersed. Theword 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 aggregatesknown as fullerene water suspensions), showed a strong tendency to aggregate in weakelectrolyte 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 ionicstrength on natural colloid aggregation was noted earlier. The n-C60 aggregateseventually settle out of suspension, sorb to particles or become otherwise immobilisedon surfaces.

Table 3. Aggregation data for manufactured nanomaterials in water (adapted from Boxall et al., 2007)

Nanomaterial Water type Aggregate sizerange, nm

Comments References

Fullerenes Freshwaterculturemedium

25-500 (mean 75) This is with a THF-based preparation.Smaller mean sizeusing sonication

Lyon et al., 2006

TiO2 Freshwater 177-810 (mean330)

From an initial sizeof 66 nm

Adams et al., 2006

SiO2 Freshwater 135-510 (mean205)


Adams et al., 2006

ZnO Freshwater 420-640 (mean480)


Adams et al., 2006

Zerovalentiron

Freshwater 1000 Rapidly aggregate Mondal et al.,2004; Schrick etal., 2004


Figure 7. Electron micrographs illustrating aggregation of zinc oxide nanoparticles from dispersion of aZnO nanopowder (nominally 30 nm) in a freshwater algal medium, pH 7.5

Zerovalent iron nanoparticles in water grow rapidly to micron sizes or more, andquickly 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, forexample, that aggregation of gold (and iron oxide) nanoparticles was minimised at lowpH. While such studies assist in understanding aggregation behaviour, they are of littlevalue 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 naturalwaters, their applicability may be problematic.

A number of papers have documented oxidation/reduction reactions of fullerenes, andthe potential for oxidation (hydroxylation) mediated by fungal enzymes has beensuggested (Wiesner et al., 2006). No such biotransformations have, however, yet beenobserved.

6.2.2 Nanoparticle Solubility

With respect to solubility, the Gibbs-Thompson effect predicts that nanoparticles with asmaller radius of curvature are energetically unfavourable and subject to preferentialdissolution, 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 Ostwaldripening, where with time, the rapid initial dilution and supersaturation solubility isreduced by the growth of larger particles with lower solubility. The overall process isone of destabilisation of nanoparticles in solution.

100 nm1 µm


These phenomena raise questions about the overall stability of nanoparticles in aquaticenvironments and highlight the need for measurements of both particle size andsolubility to reliably assess the fate of nanomaterials.

Most metal-based nanoparticles are hydrophilic and have a finite but often lowsolubility. In many cases, this is not measured, but since the soluble ionic metal fractionis the most toxic to aquatic biota, it is desirable that the extent of this solubility bedetermined. As a case in point, Franklin et al. (2007), investigating the biologicalimpacts of zinc oxide nanoparticles, found that despite a common belief that zinc oxidewas ‘insoluble’, nanoparticulate ZnO rapidly dissolved to the extent of 6 mg/L ofdissolved (dialyzable) zinc within 6 h and 16 mg/L in 72 h in a buffered pH 7.5 algalmedium. This was a concentration well in excess of the 5 mg Zn/L that would be toxicto most aquatic biota. By contrast, in similar experiments with nanoparticulate ceriumoxide, a very low solubility (ng/L) was observed, and so the effects of nanoparticleversus macroparticle toxicity could be readily investigated (Franklin, unpublishedresults). Greater toxicity to algae was observed for nanoparticulate CeO2 compared to itsmacroparticulate equivalent.

Figure 8. Illustration of the solubility of amorphous silica as a function of radius of curvature (adaptedfrom Bjorn et al., 2006)

The toxicity of a range of metal nanoparticles to a range of aquatic organisms wasinvestigated by Griffitt et al. (2008). Toxicity was observed for silver and copper with48-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 thetoxicity of nickel nanoparticles.

Semiconductor quantum dots based on cadmium selenide have been shown to releaseionic cadmium as a result of selenide oxidation (Derfus et al., 2004). Solutions of 250mg/L yielded as high as 80 mg Cd/L. The concentration of cadmium directly correlatedwith cytotoxic effects to primary hepatocytes isolated from rats and grown in vitro.CdSe/ZnS nanocrystals released a factor of 10 less cadmium, however onlypolyethylene-silane coatings were effective in preventing release (Kirchner et al., 2005).In the case of environmental release, such concentrations of cadmium would readilyexceed water quality guidelines (ANZECC/ARMCANZ, 2000) with severeconsequences for ecosystem health.

Carbon-based nanoparticles are typically lipophilic and are virtually insoluble in naturalwaters. 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 nanotubeformulations, and derivatives have been prepared with appreciable water solubility.Sayes et al. (2004) showed that cytotoxicity to human liver carcinoma cells wasinversely related to the solubility of fullerene derivatives, largely as a consequence ofthe reduced ability to generate oxygen free radicals that are the cause of cytotoxiceffects via lipid peroxidation.

It is important to recognise that the term ‘solubility’ has been loosely used by someauthors, especially in relation to carbon-based nanomaterials, often meaning formingstabilised suspensions as distinct from truly dissolving as was the case with metaloxides 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. Theseformulations may also result in different solubility characteristics. Carbon nanotubesare extremely hydrophobic and subject to high van der Vaal’s forces along the lengthaxis, with a tendency to aggregate. To disperse CNTs in aqueous solution, a range ofchemical additives have been used including surfactants (sodium dodecylsulfate,sodium dodecylbenzene sulfonate, Triton X-100) and polymers, acting either stericallyor electrostatically (Brant et al., 2005). The effect of these dispersant additives isusually to sterically stabilise the nanomaterials, by physically hindering theiraggregation (Handy et al., 2008; Saleh et al., 2008). Terashima and Nagao (2007)showed that the surfactant Triton X-100 and natural humic substances enhance thesolubility of C60 nanoparticles by 8-540 times, while also decreasing the rate ofaggregation (Chen and Elimelech, 2007).

In many cases the effect of additives on solubility and aggregation in commercialnanomaterial formulations is unknown. In the case of zinc oxide formulations, forexample, a relevant concern might be whether the equilibrium water solubility ofnanoparticulate zinc oxide in sunscreens is different to that seen for the raw


nanoparticles by Franklin et al. (2007). Such questions have important implications forrisk assessment in aquatic systems.

Nanomaterials often contain impurities, for example, carbon nanotubes have beenreported 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 theweight 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 evenhigher in those prepared by chemical vapour deposition. These impurities can affect thesurface charge, reactivity, transport and ecotoxicology of the SWCNTs.

The presence of impurities was found to be responsible for oxidative stress damage torat 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 tolarge mouth bass (Brant et al., 2005). While these impurities are not expected tosignificantly affect nanoparticle fate, they raise interesting questions with respect to theeffects 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 validconcern.

6.2.4 Fate in Natural Water Systems

While there are data from laboratory studies on the behaviour of selected nanomaterialsin water, the behaviour is likely to differ in natural waters, where there is a possibility ofinteraction with natural colloids including dissolved (and particulate) organic matter(NOM). The importance of colloids cannot be underestimated. In freshwaters forexample, colloidal organic matter concentrations lie in the range 1-10 mg/L compared tothe concentrations that have been predicted for manufactured nanoparticles of 1-100µg/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 SuwanneeRiver humic acid greatly enhanced the dispersion of multi-walled carbon nanotubes inMilli-Q water, and that the same effects were also seen in suspensions in SuwanneeRiver water samples. The dispersion was greater than that observed with sodiumdodecylsulfate. The exact mechanism of the enhanced dispersion is likely to againinvolve both steric and electrostatic components, as was seen for natural colloids.Similar stabilisation of iron oxide nanoparticles by humic acids has also beendemonstrated (Tipping and Higgins, 1982; Baalousha et al., 2008).

By contrast, it has been suggested that natural fibrillar colloids are likely to increaseaggregation because of different binding characteristics, compared to the chargestabilisation mechanism of humic substances (Buffle et al., 1998).

The findings to date suggest that in natural water systems, nanoparticles may have agreater 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 removalmechanism that could enhance transport to and accumulation in bottom sediments.

Given these many uncertainties, site-specific fate studies are recommended that useactual nanomaterial formulations in a variety of natural waters (fresh and estuarine).

Where nanoparticles are released with wastewaters, it has been suggested that thepresence of household or industrial detergents would result in the disaggregation ofnanoparticles (Fernandes et al., 2006). In a related study of cerium oxide nanoparticlesin a model wastewater treatment system, Limbach et al. (2008) found that a small butsignificant fraction (6%) avoided aggregation and was released in the effluent (at 2-5mg/L concentrations) largely as a result of stabilisation in the presence of proteinbreakdown products and surfactants in the wastewater changing the zeta potential.These examples highlight the impact that surface modifications from wastewatercomponents 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 onthe fate of nanoparticle silver released into water from commercially available sockfabrics. 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 treatmentprocess they showed that all of the silver was removable to the sludge, raising concernsabout the impacts of application of sludge to land.

The behaviour of emulsions in natural waters has been poorly studied. In a report onacrylic latex, NICNAS (2000) noted that ‘the fate of the aqueous residues released to thesewer system is less predictable as the notified polymer may remain in the aqueousphase as an emulsion at low concentrations’. In addition, ‘all solid residues will remainassociated with the soil and sediment due to the high molecular weight and the stabilityof the cured paint matrix’.

6.2.5 Nanoparticles as Vectors for Contaminant Transport

While so far we have considered manufactured nanoparticles as potential sources oftoxic effects in the environment, as noted earlier with respect to colloids, nanoparticlesare excellent binding sites for other soluble contaminants and therefore have thepotential to act as vectors for the delivery of these contaminants. Again, this will be afunction of surface properties of the particular nanomaterial formulations.

A recent publication by Hu et al. (2008) showed that aqueous suspensions of fullerenewere able to effectively sorb polycyclic aromatic hydrocarbons (PAHs), a process thatwas further enhanced by the addition of humic acids. This predictable behaviourindicates that nanomaterials can affect the fate of hydrophobic organic contaminants innatural waters.


In soils, groundwaters, rivers and lakes, natural colloids have been shown to play animportant role in trace metal retention and transport (Kretzschmer and Schafer, 2005).Similar binding capacities exist for manufactured nanoparticles. Secondary toxicityeffects from these adsorbed contaminants will need to be considered in any toxicitystudies of nanoparticles.

6.3 Fate of Manufactured Nanomaterials in TerrestrialSystems

There is currently very little information available with which to assess theenvironmental risk of manufactured nanoparticles to terrestrial ecosystems. The keyphysico-chemical properties of nanoparticles described above are also likely to play amajor role in the fate, transformation, and environmental effects in soils. Soils differfrom 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 detectionof the manufactured nanoparticles in the presence of natural nanoparticles, which areubiquitous in soil.

6.3.1 Key Pathways

A number of key processes are likely to affect the fate and bioavailability ofnanoparticles in the soil environment (Figure 9).

Nanoparticles have high surface reactivity and, depending on surface charge andcoatings, their adhesion to reactive soil surfaces may be strong –“partition coefficients”for nanoparticulate contaminants in soil have yet to be published. Data from transportstudies of soil colloids however indicate that surface coatings on the nanoparticles areimportant 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 fornanoparticles used in groundwater remediation (Hydutsky et al,. 2007). As yet, thereare few data on transport of nanoparticles through soils, and hence characterisation ofnanoparticle 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 orderas listed (Lecoanet et al., 2004). Another recent sand column study demonstrated theimportance of surface coatings in the transport of zerovalent iron nanoparticles (Saleh etal., 2008). Similar studies in soils are now required.


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

Figure 9. Key processes in soil relating to transformation and potential risk from manufacturednanoparticulate particles

6.3.2 Behaviour of Natural Colloids in Soils

Naturally-present colloids and macromolecules in soils are similar to those found infreshwater systems, and nanoparticulate and microparticulate clays, organic mater, ironoxides and other minerals play an important role in biogeochemical processes. Soilcolloids 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 surfacewaters in Australia, and studies have examined the factors responsible for colloidgeneration 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 colloidsin relation to soil physical and chemical properties (for a review see Rengasamy andOlsson, 1991). Aggregation of colloids in soil is a function of surface charge, ionicstrength, particle size and chemical composition of the soil pore water and exchangeableions held on the surface of colloids. Systems dominated by sodium and with low ionicstrengths are likely to have dispersion of colloids, while those dominated by calciumand high ionic strengths will tend to aggregate. Recent evidence confirms thatmanufactured 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 mineralsand organic matter.

6.3.3 Behaviour of Manufactured Nanoparticles in Soils

Of the pathways identified in Figure 9, the most important properties that will controlnanoparticles fate in soils are likely to be dissolution, aggregation and partitioningbetween solution and solid phases.

Nanoparticle Solubility

Dissolution of nanoparticles in aqueous media has already been covered in Section 6.2.2above. A key difference in soils is the large surface area and exchange capacity forcations and anions that can promote dissolution of compounds through acting as a sinkfor dissolution products, and providing protons to enhance dissolution of compoundswith a pH-dependent solubility. To date, no studies have examined the rate or extent ofdissolution of nanoparticulate materials in soils in relation to their bulk counterparts.

Aggregation

There are virtually no studies which have examined this topic for manufacturednanoparticles in soils, but inferences from the behaviour of natural colloids can be made(Section 4.2 above). Aggregation behaviour of nanoparticles in aquatic systems hasbeen covered in Section 4.2, and the same processes would be active in soils, except wecan speculate that the aggregation of nanoparticles in soil may be greater due to thehigher ionic strength of soil pore waters compared to most surface water systems(streams and dams). Aggregation in soils also leads to particle entrapment in poresthrough 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 thatthe high surface area and charge of many hydrophilic manufactured nanoparticles willcause a strong binding to the predominantly negatively charged surfaces of soil mineralsand organic matter (Li et al. 2008), depending on the nature of the charge. Netpositively charged particles will be retained strongly, while those with net negativecharge will be highly mobile in most soils (Saleh et al., 2008).

Where nanoparticles are hydrophobic, retention to organic matter surfaces in soil mayinhibit mobility and availability to organisms.

The characteristics of the surface “functional” coatings used in nanoparticlemanufacture may be very important in explaining (and predicting) fate in soil, as it isthese 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 riskassessments for a wide range of inorganic and organic contaminants in terrestrialsystems, this characteristic is a key property requiring evaluation.


7. ECOLOGICAL RISK ASSESSMENT OF MANUFACTUREDNANOPARTICLES

Regulators worldwide are seeking to undertake ecological risk assessments ofmanufactured nanoparticles to determine the significance of any impacts associated withtheir manufacture and use. The ecological risk assessment framework for contaminantsin the environment, developed by the USEPA and adopted in Australia, has thefollowing components:

Problem formulation

Exposure assessment – Chemical assessment taking into account contaminantfate (predicted environmental concentrations – PECs)

Effects assessment – Measurement of toxicity, bioaccumulation, effects onecology (predicted no effect concentrations – PNECs)

Risk characterisation (PEC/PNEC)

As discussed by Owen and Handy (2007), the issue of problem formulation is a criticalone. The initial anxiety that nanomaterials might represent the current equivalent ofgenetically modified foods in terms of its environmental danger (Dowling, 2005)appears to have now passed. These concerns were heightened by the findings thatfullerenes 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 thebasic issue of whether nanosized materials pose a greater hazard to biota than theequivalent macrosized materials. While there is good evidence for altered behaviourwith smaller size, only a handful of studies have demonstrated that this translates intogreater toxicity. The risk assessment needs to show connectivity between the source,the pathway, and the receptor. In most instances in water and soil ecosystems, theevidence of this connectivity has been indirect or absent.

For industrial chemicals, a manual providing guidance on ecological risk assessmentwas recently released by the Department of the Environment and Water Resources (nowDepartment of the Environment, Water, Heritage and the Arts, DEWHA) (DEW, 2007).This manual specifically discussed data requirements, data evaluation, environmentalexposure 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, watersolubility, hydrolysis as a function of pH, octanol/water partition coefficient, adsorptionbehaviour in soils, acid dissociation constant, and environmental fate data, especially onbiodegradation and bioaccumulation. For effects assessment, toxicity tests must beundertaken using a fish acute test, Daphnia immobilisation and reproduction tests, analgal growth inhibition test, and measures of biodegradability and bioaccumulation. Asthe following pages will show, the majority of these requirements could not currently bemet for manufactured nanoparticles. This means that the determination of both PECs


and PNECs will not be possible as a prerequisite to assessing the potentialenvironmental hazard of manufactured nanoparticles in soil, water and sedimentcompartments. The current state of knowledge in these areas is reviewed in thefollowing pages of this report.

7.1 Polymeric Nanoparticles as a Separate Class

Some authors have drawn the distinction between polymers such as dendrimers,fullerenes and carbon nanotubes whose size is determined by their molecular weight andother particles where size is a function of their degree of aggregation. There is potentialfor the size of both nanoparticle types to affect their interactions with aquatic biota, butto date few studies have investigated this. We see no reason to consider polymers aswarranting separate regulatory consideration from other nanoparticle types.

It is clear that assessment of hazard of either type on the basis of intrinsic chemicalproperties is inappropriate and there must be some consideration of size, be it molecularweight or other measures of size. The premise that nanoparticulate size fractions aremore toxic than larger size fractions needs to be tested for all nanoparticle classes withrespect 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 formsof a contaminant in the environment that, with a consideration of fate and exposureduration, can be linked to effects on target organisms. It will be important therefore thatmeasurements reflect the concentrations and physical and chemical properties of thenanoparticles in the field that are truly representative of exposure. In assessingindustrial chemicals, the DEW manual (DEW, 2007) lists fate, partitioning behaviour,and persistence as important parameters. These need to be combined with concentrationdata in estimating likely exposure.

For nanomaterials, since it has been demonstrated that size is a critical parameter, anymeasurement of concentration must be accompanied by data on the distribution ofparticle sizes in the test water taking into account any aggregation that might occurwithin the life cycles of the test organisms.

The requirement and the current status of methods for nanoparticle analysis andcharacterisation 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 whenparticles were other than spherical, and other measures including aspect ratio (ratio oftheir longer dimension to their shorter dimension) were also of value. They believedthat in addition to particle size distributions, measures of surface area were alsoimportant, 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 electrostaticrepulsive forces.

An interesting issue is the extent to which nanoparticle size distributions reach steadystate, 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 whenstudying the effects of nanoparticulate titanium dioxide on rainbow trout. Hassellov etal. (2008) recommended that monitoring be undertaken over the duration of any studiesto 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 areaand charge. Where possible, measurements should be of the state of the nanoparticles inthe particular medium (soil, sediment, water), rather than an assumption based ondilution of a starting material. Even the measurement of concentration poses issues, aswe 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 thannanoparticulate zinc oxide. In the case of n-C60, UV absorbance was used to measureconcentrations (Oberdorster et al., 2004), while for carbon nanotubes, light scatteringtechniques were used to correlate with concentration (Smith et al., 2007).

The bioavailable fraction can however include a dissolved, soluble fraction rather than ananoparticulate fraction, so ideally some measurement that discriminates this fraction isrequired. Standard 0.45-µm membrane filtration will not retain most nanoparticles, so aseparation technique is required. Ultracentrifugation, size-based chromatographicseparations, ultrafiltration and dialysis are all appropriate, although the last two areprobably the preferred methods of separation. In soils, there is the additionalcomplication that any nanoparticulate material that dissolves will interact with the soilsolid phase, and some assessment of this pool may also be required to assessbioavailability in addition to characterisation of the material in soil pore water.

For measuring particle size distributions, electron microscopy (EM) and dynamic lightscattering (DLS) are the most commonly used techniques. Both have advantages anddisadvantages (Bootz et al., 2004). EM gives the most direct information on the sizedistribution and shapes of particles, however, there is concern about artefacts introducedby the sample preparation step. With DLS, the presence of small amounts of largeaggregates can affect the distribution of a main component of a smaller size, with resultsbeing misleading where the samples have a broad size distribution. More detailedinformation on specific surface area, surface charge and zeta potential can be obtainedby a variety of techniques, but these are research techniques that are not likely tocontribute 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 fieldflow 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 beanalysed by inductively coupled plasma mass spectrometry (ICPMS). Stolpe et al.(2005) have described the application of high resolution ICPMS coupled to FlFFF tostudy metals in (natural) nanoparticulate colloids. The FFF technique has been wellestablished, but is not that easily mastered, and is in use in only a handful of laboratoriesworldwide. The universal interest in nanomaterials might lead to a wider acceptance.

Single particle ICPMS analysis has recently been applied to the detection of goldcolloids in water (Degueldre et al., 2006). The use of new generation ICPMSapproaches for analysing individual nanoparticles show considerable promise (Stolpe etal., 2005), but it may be some time before they can be applied to routine environmentalmonitoring of manufactured nanoparticles.

Similarly the use of novel techniques such as liquid chromatography coupled to nuclearmagnetic resonance spectrometry may hold promise for the analysis of carbon-basednanoparticles.

The analysis of manufactured nanoparticles in natural systems can be complicated bythe 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 notbe able to discriminate between nanoparticles and natural colloids unless linked to somenanoparticle element-specific analyses such as ICPMS. The solution is to use singleparticle confirmatory analyses, such as EM, with energy-dispersive x-ray fluorescence(EDX).

To date there have been few measurements of manufactured nanoparticles in naturalwaters or soils because of the extreme difficulty in detecting environmentalconcentrations. Details of techniques applied to environmental nanoparticles in aquaticsystems have been discussed by Wiggington et al. (2007).

The area of nanoparticle measurement is one that is being pursued internationally by anumber of agencies. In particular, there is a need to develop standard methods ofanalysis, including methods for sample preparation that can be used to characterisenanoparticles. As part of this exercise, the development of standard reference materialsthat can be used for method quality assurance and quality control will be essential. InAustralia, 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 ona few critical tests recommended by the Organisation for Economic Cooperation andDevelopment (OECD) (OECD, 2007), including those for water solubility,adsorption/desorption, water/oil partition coefficient, hydrolysis, surface tension and fatsolubility. The applicability of each of these tests to nanomaterials is generallyinappropriate, and the methods will need to be considerably altered to adequately caterfor nanomaterials.


The test for water solubility (No. 105) (OECD, 2007) uses either a microcolumnseparation or a flask dissolution. Since the tests were not designed for use withcolloidal or nanosized particles, the separation of these from the ‘soluble’ fraction willbe critical. The test method indicates that the presence of colloids in the microcolumneffluent invalidates the test. In studies of zinc oxide solubility, Franklin et al. (2008)used dialysis to separate soluble zinc. Such procedures will be required as a finish tothe OECD test. The same applies to Test No. 120, for the solution/extraction behaviourof polymers in water.

The tests looking at adsorption/desorption onto soils need to be relevant to the likelyenvironmental concentrations. Test No. 106 uses a soluble chemical fraction, however,for a nanomaterial suspension, this would not be appropriate. Test No. 121 determinesthe adsorbed fraction by HPLC. Determining whether the nanoparticles are retained byfiltration rather than adsorption will be problematic.

The octanol/water partitioning tests (Nos 107 and 123) are designed to measure theequilibrium partitioning of a ‘dissolved’ substance between the two solvents, as distinctfrom ‘solubility in octanol’ which is what will be obtained using nanoparticles. Even ifit were meaningful, the physical application of this test is likely to be seriously impairedby clumping of nanoparticles at the solvent interface.

The hydrolysis test (No. 111) looks at hydrolysis in the range pH 4-9. Withnanomaterials, the result would test both dissolution and hydrolysis as a function of pH.

Surface tension (No. 115) is inappropriate for an insoluble chemical in nanoparticulateform, however, the assessment of fat solubility (No 116) is a potentially useful measurein relation to biological uptake.

OECD has an active interest in nanomaterials, and has a working group consideringappropriate 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 tonanoparticles. As already discussed, there have been preliminary approaches topredictive modelling of the suspension stability and kinetics of aggregation ofnanoparticles, however their applicability to real systems is, as yet, untested (Mackay etal., 2006).

In an attempt to model likely concentrations of manufactured nanoparticles that mightbe found in the environment, Boxall et al. (2007) used a series of simple algorithms topredict 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 frombioremediation;

(ii) inputs from spray drift following use of agrochemicals;


(iii) runoff from contaminated soils;

(iv) aerial deposition; and

(v) emissions from wastewater treatment plants.

For soils, routes comprised:

(i) the application of remediation technologies;

(ii) the application of plant protection products;

(iii) the excretion of nanomedicines used in veterinary products;

(iv) aerial deposition; and

(v) the application of sewage sludge as a fertiliser.

They focussed mainly on cosmetics, personal care products and paint, and thenanoparticle concentrations that they contained (based on limited European data). Threehypothetical scenarios were modelled, where 10, 50 and 100% of a product typecontained the manufactured nanoparticle. Predicted concentrations for the 10% scenarioare shown in Table 4. Despite all of the uncertainties, the concentrations can becompared to the toxic concentrations where these are known, to see whether these are inthe same range or not.

Table 5 shows the comparison of exposure data with known toxicity data, indicatingthat the predicted environmental concentrations are orders of magnitude below thoseknown 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 accidentalreleases. The results nevertheless give regulatory agencies some reassurance, especiallysince the assumptions in estimations are conservative.

The challenge for modellers in the derivation of appropriate PECs is to be able to obtainreliable estimates of the mass flow of nanomaterials to different compartments of theenvironment. A good example of a life cycle assessment approach to this is shown inFigure 10 (from Blaser et al., 2008). This example has been used for silver derivedfrom nanoparticulate biocidal plastics and textiles, but the approach has genericapplication. In deducing mass flows, estimates of total product usage and estimated (ormeasured) release rates must be obtained. These data are then related to the time ofexposure. Knowledge of the behaviour of silver in the aquatic environment (colloidalforms, attachment to particles, etc.) is used in coupled river fate models to predictsediment/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

Aluminium oxide Paint 0.002 0.01

Cerium dioxide Scratch resistant coatings, catalysts <0.0001 0.01

Fullerenes Anti-inflammatory cream, eyeliner, facepowder, 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 Boxallet al., 2007)

Predicted inwater, µg/L

Toxicity data, µg/L Other endpoints

InvertebrateEC50

FishLC50

AlgaeEC50

n-C60 0.31 >35,000 >>5000 - Effects on invertebrate growth at260 µg/L; bacterial growthaffected at 40µg/L; bacterialphospholipids affected at 10 µg/L

TiO2 24.5 >100,000 >100,000 16,000 Effects on invertebrate growth at2000 µg/L; bacterial growthaffected at 100,000 µg/L

SiO2 0.0007 - - No effect on bacterial growth at500,000 µg/L

ZnO 76 - - No effect on bacterial growth at100,000 µg/L


Figure 10. Framework for deriving mass flow data for silver flows from nano-functionalised biocidalplastics and textiles (from Blaser et al., 2008). Arrows represent silver flows; dashed lines indicatedifferent environmental spheres. TWT=thermal waste treatment; STP=sewage treatment plant.

The model predictions can be verified by comparison with measured data from differentaquatic environments.

Mueller and Nowack (2008) have followed a similar approach in the determination ofthe expected exposure concentrations in air, soil and water for nanoparticulate silver andtitanium dioxide and for CNTs. Literature production data are used to determine theweighted concentrations of nanomaterials from each product type. Release ofcontaminants and their transfer between the various compartments in the model are thendetermined using literature-derivations or best estimates of transfer coefficients. Inmuch the same way as that used by Boxall et al. (2007), PECs were derived forparticular environmental compartments, however, the number of product categories wasextended beyond the personal care and cosmetic products, to include all possible uses,e.g. for silver, the categories were textiles, cosmetics, metal products, sprays andcleaning agents, plastics, and paints. The findings are shown in Table 6.

The results were then compared with available toxicity data. No data were available forsoil toxicity. The EC50 value (concentration causing a 50% effect) used for silvertoxicity was 20-40 mg/L, but this was from bacterial toxicity testing (E. coli andBacillus subtilis), and so are not necessarily applicable. The authors noted that, forionic 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 hazardquotients (PEC/PNEC) indicate a potential concern for TiO2, compared to theconclusions of Boxall et al. (2007) discussed above, but this may be a function of theapplication 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, waterand soil (RE: realistic scenario; HE: high emission scenario) (from Mueller and Nowack, 2008)


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 affectedby 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 affectedby wastewater

0.2 0.5 >180 > 3900 naa na

aNot available

9. ECOTOXICOLOGY OF NANOPARTICLES

9.1 Ecotoxicity and Nanoparticle Dose Metrics

In examining the ecotoxicity of nanoparticles to biological organisms, a critical questionis the determination of what influences the dose response. In traditional ecotoxicologywith soluble species, concentration is the dose measure, and specifically, thebioavailable concentration, which may be some sub-set of the total concentration.When dealing with nanoparticles, the concentration or mass metric may not apply, andalternative considerations may involve particle number, surface area (shape), particlecomposition, or surface reactivity.

In defining the applicable dose metric, some understanding of the mechanism of toxicityof nanoparticles is required. Thus toxicity could be exerted by soluble speciesdissociating from nanoparticles at a cell surface and crossing the cell membrane, or bydisruption of cell function by blockage of surface sites. If the nanoparticle is aheterogeneous 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 thenanoparticles that are capable of free radical generation.

In studies of human toxicology of nanoparticles, there has been some debate about theappropriate dose metric. Oberdorster et al. (2005) showed that surface area accountedfor differences in lung inflammatory effects of nanoparticulate TiO2 to rats and mice farbetter than any mass considerations. Duffin et al. (2002) reached similar conclusionsfor quartz nanoparticles, although noting the importance of surface reactivity.Wittmaack (2007) disputed this interpretation, suggesting that particle number provided


a better fit for differently prepared carbon nanoparticles, although the interpretation wascomplicated by the possibility that aggregated particles might disaggregate on contactwith the lung. The relevance of these studies with atmospheric nanoparticles to toxicityin aquatic or soil systems is, however, questionable.

Particle morphology may also be an important metric (Buzea et al., 2007). Particles canbe 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 aspectratio. In pulmonary toxicology, particles with a high aspect ratio have been shown to bemore toxic (Inoue et al., 2006). The importance of aspect ratio in aquatic or terrestrialtoxicity is unknown.

No such definitive studies appear to have been undertaken for ecotoxicologicalreceptors, although there is clear evidence of greater toxicity of nanoparticulate versuslesser 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 sincebioavailability will be dependent upon the physical properties, it will be necessary toqualify all concentration data. Size is the next most critical parameter, since indicationsare that when the nanoparticulate range is exceeded, properties approach those of theparent macroparticles. Included in the size estimation should be a verification of thefraction that is not in true solution, so that any observed effects are related only to theparticulate forms. It should be noted that the key environmental issue for any riskassessment is to derive a no-effects concentration in the site-specific medium. Toxicitydetermined in synthetic media might greatly over- (or under-) estimate toxicity becauseof modification of bioavailability in the presence of colloids or other constituents.

The parallel in aquatic ecotoxicology to the consideration of site specific modificationsto water quality guidelines is a useful one (e.g. ANZECC/ARMCANZ, 2000). Thetoxicity 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 thisstage of our knowledge, other physical and chemical measurements are probablysuperfluous, given the already existing uncertainties in the measurements that are beingmade.

A major practical issue with toxicity testing of nanomaterials is the dispersion ofnanoparticles in the test solutions. In aquatic toxicity testing, the contaminant is usuallyin true solution and homogeneously distributed throughout the sample. Any attempts touse artificial dispersants or sonication are likely to affect the degree of aggregation fromits natural state and so stirring of the sample is the only acceptable means of maintainingthe nanoparticles in any way close to a dispersed state, acknowledging that prolongedstirring may also break up nanoparticles. Intermittent stirring might be an option.

A framework for nanoparticle toxicity assessment in waters based on the abovediscussion is given in Table 8. A similar approach could be devised for testing in soils.


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 anappropriate 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 totalnanomaterial 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 TriggerValues

1. Repeat the above approach using either a synthetic or site water sample asappropriate.

2. Compare result with trigger value, noting compatibility of particle sizedistribution.

Crane and Handy (2007) in a recent review of methods for characterising theecotoxicological hazard of nanomaterials suggested that rapid tests that identifiedspecific modes of toxicity, e.g. genotoxicity, immunotoxicity or oxidative stress assaysmight 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 onnanomaterials, 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 oftoxicity 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: oxygenions, peroxides and free radicals) (Oberdorster et al., 2005, Nel et al., 2006).

Physical damage to cell membranes is also possible as a consequence of the abrasivenature of some nanoparticles leading to toxicity (Stoimenov et al., 2002). Adhesion ofnanoparticles to the cell surface and dissociation of soluble toxic species can alsoprovide 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 fromnanoparticulate ZnO because of the unexpectedly high solubility of ZnO. Theysubsequently demonstrated enhanced toxicity of CeO2 nanoparticles compared to bulkCeO2 (Franklin et al., unpublished results), suggesting enhanced uptake.

For aquatic biota, nanoparticle uptake and potential toxicity will be dependent on thetype of organism, its trophic level and whether it is uni- or multicellular. Withunicellular organisms, the issue of whether nanoparticles can cross cell membranesdirectly or via endocytosis is still a major question. For eukaryotic organisms, mostinternalisation of nanoparticles will occur via endocytosis (Moore, 2006; Nowack andBucheli 2007), i.e. with the cell membrane enclosing the nanoparticles leading to theirdeposition in the cytoplasm and association with intracellular organelles, withoutdirectly passing through the cell membrane.

For higher organisms, uptake across the gill and other external surface epithelia is alsopossible and interactions with aquatic plants may include adsorption onto the rootsurface, 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 apossibility for many organisms. Water fleas (Dapnia magna) rapidly ingested lipid-coated nanotubes via normal feeding behaviour, metabolizing the lipid coating as a foodsource (Roberts et al., 2007). The toxic impact in many instances will depend on theability 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-lethalconcentrations for 10-days, demonstrated an increase in oxidative stress markers andionoregulatory disturbance (Smith et al., 2007). More recently, direct evidence for adietary pathway of nanoparticle uptake has been demonstrated for uptake of quantumdots 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 aresummarized in Table 9.

Bacterial toxicity

Many of the toxicity assessments of nanomaterials have focussed on bacteria, largelyundertaken using traditional growth media under optimum conditions. These data havebeen 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 thatproperty is the reason for its extensive usage. Similar antimicrobial activity is shown bytitanium dioxide (Wolfrum et al., 2002). More recent studies have demonstrated strongantimicrobial activity of SWCNTs (Kang et al., 2007).

While these studies have been useful in investigating mechanisms of toxicity andrelative 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) bacterialdata are not used in species sensitivity distributions to determine safe concentrations ofnanomaterials in waters (DEW, 2007).

As already discussed in Section 5, the potential for impact on sewage bacteria is aseparate question to protecting organisms in natural waters.

Algal toxicity

Limited data are available for algal toxicity. The response to TiO2 is not particularlysensitive (Hund-Rinke and Simon, 2006; Warheit et al., 2007) and that to ZnO is aresponse to soluble zinc (Franklin et al., 2007).

Invertebrate toxicity

The freshwater crustacean Daphnia magna has been the most used invertebrate speciesfor nanomaterial toxicity testing. Daphnia were quite sensitive to n-C60 prepared bytetrahydrofuran (THF) extraction (Zhu et al., 2006; Lovern and Klaper, 2006). It isimportant to note that for these fullerenes, two preparation methods were followed, oneusing sonication of fullerenes in water for 30 minutes to disperse the nanoparticles andthe second using the evaporation of THF from a THF extract added to water. The latterwere consistently more toxic to all organisms tested, and the question remains as towhether the additional toxicity was due to THF, although these tests used THF onlycontrols. 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 thanalgae or daphnids. This was not necessarily the case with nanomaterials, and may beindicative of a different mechanism of toxicity, e.g. gill clogging, that would not occurwith dissolved contaminants.

Toxicity of nanoparticulate silver to zebrafish embryos has been demonstrated by Lee etal., (2007). In this study, the only one to date of nanoparticulate silver toxicity toaquatic biota, the endpoints were deformities and abnormalities in the embryos. NoEC50 values were quoted, but from the graphs were estimated to be in the range 10-20ng/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) reporton acrylic latex indicates that no toxicity data are available. They are generally believedto have low toxicity.


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


n-C60 water-

solubilised,

Nominally 10-200 Synthetic hard water Freshwatercrustacea

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 Meiobenthiccopepods

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

Templeton et al., 2006

SWCNT asprepared

? Seawater Meiobenthiccopepods

Amphiascus tenuiremis No effect at 1.6 mg/L; effects at10 mg/L.

Templeton et al., 2006

SWCNT ? Freshwater with up to 0.15 mg/LSDS

Rainbow trout Oncorhynchus mykiss Effects on ventilation rate, gillpathologies and gill mucussecretion 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 waterOECD 201 protocol

Algae Desmodesmussubspicatus

Chlorophyll fluorescence72-h EC50 44 mg/L small; nodose response large

Hund-Rinke and Simon,2006

TiO2 Average 140 Moderately hard water

OECD 201 protocol

Algae Pseudokirchneriellasubcapitata

Chlorophyll fluorescence

72h EC50 16-21 mg/L

Warheit et al., 2007

TiO2 Nominal 25(small); 100(large)

Moderately hard waterOECD 202 protocol

Water flea Daphnia magna No concentration-effect curveobserved up to 3 m/L

Hund-Rinke and Simon,2006

TiO2 THFdispersed;sonicated

30 THF; 100-500sonicated

Moderately hard freshwaterUSEPA protocol

Water flea Daphnia magna 48-h LC50 THF 5.5 mg/L;Sonicated >500 mg/L

Lovern and Klaper,2006

TiO2 THFdispersed;sonicated

30 THF; 100-500sonicated

Moderately hard freshwaterUSEPA protocol

Water flea Daphnia magna No significant behaviouralchanges LOEC 2.0 mg/L

Lovern et al., 2007


TiO2Average 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-dayexposure up to 1.0 mg/L.Sub-lethal effects including gilldamage, observed.

Federici et al., 2007

TiO2140 Moderately hard water

OECD 201 protocolRainbow 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 dayexposure to 10 mg/L.

Sun et al., 2007

TiO2Nominal 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 todissolved 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


shell

Quantum dots;

Cd/Se or Cd/Te

core with ZnS

shell

Estimated 10-25 Moderately hard water Algae Pseudokirchneriellasubcapitata

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


9.2.3 Developing Appropriate Guidelines for Nanomaterials in Waters

The available toxicity data are insufficient to develop reliable guidelines for mostnanomaterials in waters, however, it is instructive to attempt to derive low reliabilityguidelines for the nanomaterials for which we have the most data. For the data fromTable 6 for n-C60 and TiO2, chronic NOEC values were obtained using a factor of 10 onacute LC50 values or on EC50 values from acute endpoints (Table 10). FollowingANZECC/ARMCANZ (2000) guidelines, data would be required for an alga, aninvertebrate 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 7.9µg/L would be derived for the water-solubilised n-C60. A value for THF-extract n-C60 ismore problematic and clearly <0.5 µg/L. The OECD approach would use a factor of1000 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 Estimatedchronic NOEC,

mg/L

n-C60 Watersolubilisedby sonication

Daphnia magna 48-h LC50 7.9 0.79

n-C60 Watersolubilisedby sonication

Pimephalespromelas

No effects after48 h at 0.5

>0.05

n-C60 THF extract Daphnia magna 48-h LC50 0.8,0.46 (acuteNOEC 180 µg/L)

0.08, 0.05

n-C60 THF extract Pimephalespromelas

100 % mortalityin 6-18 h 0.5

<0.05

n-C60 THF extract Mycropterussalmoides

100% mortalityin 6-18 h 0.8

<0.08

n-C60 THF extract Danio rerio <1.5 <0.15

TiO2 No THF Desmodesmussubspicatus

72-h EC50 44 8.1

TiO2 No THF Pseudokirchneriellasubcapitata

72-h EC50 16-21

4.0

TiO2 THF extract Daphnia magna 48-h LC50 THF5.5

0.55

TiO2 No THF Daphnia magna >500 >50


With this extra conservatism, the PNEC value for both n-C60 and TiO2 are seen to benow getting closer to the PEC values estimated in the UK (Table 5), with all of theirlimitations. 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 nanomaterialsby aquatic organisms is limited and is summarised in Table 11. There is TEM evidenceof the presence, in the cytoplasm of bacterial cells (e.g. Escherichia coli, Bacillussubtillus, Staphylococcus aureus), of MgO (Makhulf et al., 2005), SWCNTs (Kang etal., 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 cellulardamage, however, intracellular uptake was only indicated for MgO (<11 nm) and silvernanoparticles (<80 nm), and for quantum dots (<5 nm).

As noted earlier, an important finding was the food chain transfer of quantum dots viaexposed algae to water fleas (Bouldin et al., 2008). The quantum dots have a CdSe coreand 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 levelswell 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 TargetOrgan

Evidence Reference

Bacteria

MgO Escherichia coli

Bacillussubtillus

Membrane TEM images confirm damageand leakage of cell contents.

Stoimenov,2002

SWCNT Escherichia coli Membrane Increased membranepermeability in cells in directcontact with SWCNT.Physical damage to themembrane and leakage of cellcontents is proposed.

Kang et al.,2007

MgO Escherichia coli

Staphylococcusaureus

Whole cell TEM shows ultrastructuralchanges on exposure to 8±1and 11±1 nm particles.Elevated intracellular Mgconfirmed.

Makhluf et al.,2005

ZnO Escherichia coli Whole cell TEM reveal electron denseareas in the cytoplasm. Noelemental analysis.

Brayner et al.,2006

Quantum dots Escherichia coli Whole cell TEM, fluorescencespectroscopy show adenine-

Kloepfer et al.,


Bacillus subtilis conjugated QDs < 5 nm areinternalised. Intracellular Cdand Se confirmed.

2005

Ag Pseudomonasaeruginosa

Whole cell Particles up to 80 nmtransporting in and out ofcells. TEM images confirmelectron dense areas in thecytoplasm.

Xu et al., 2004

Ag Escherichia coli Whole cell TEM images showingelectron dense intracellularareas. EDS elementalmapping confirms Agdistribution throughout thecell. 1 -10 nm particlesinteract preferentially withthe cell.

Morones et al.,2005

SWCNT lipid-coated

Daphnia magna Gut Rapid (45-min) ingestion andpresence of lipid-coatedSWCNT in the gut trackobserved in time-coursemicrographs.

Roberts et al.,2007

Quantum dots Ceriodaphniadubia

Gut Evidence for food chaintransfer of core metals fromquantum dot-dosed algae

Bouldin et al.,2008

Fish

Cu Danio rerio Gill Histopathological analysisrevealed damage to gilllamellae by proliferation ofepithelial cells and oedema ofgill filaments. Unclear ifeffects mediated by particleuptake.

Griffitt et al.,2007

TiO2 Oncorhynchusmykiss

Gill

Gut

Histopathological changes tothe gill and gut but fish didnot accumulate TiO2 in theinternal organs.

Federici et al.,2007

SWCNT Oncorhynchusmykiss

Gill

Gut

Histopathological changes tothe gill and gut and liver.Aggregated SWCNTsobserved in the gut lumen.

Smith et al.,2007

9.2.5 Ecological Impacts

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


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. Thenanoparticles are likely to be highly aggregated in the sediments, so any unique toxicproperties associated with nano size are likely to be absent. Very few studies havelooked at nanomaterials in sediments. For example, Kennedy et al. (2008) showed thatthe survival of several amphipods was affected by MWCNTs in whole sedimentbioassays, but at unrealistic concentrations exceeding 100 g/kg. More studies arerequired 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 ofnanoparticles to the terrestrial environment and this is seen as a key knowledge gap byregulators (US EPA , 2007). As yet, there are few reports in the peer-reviewedscientific literature of the assessment of ecotoxicity of nanoparticles to soil biota, insoils. Several reports have examined ecotoxicity to soil organisms, but the media usedhave been simple aqueous media (Brayner et al., 2006; Yang and Watts, 2005; Zheng etal., 2005; Lin and Xing, 2007) and persistence of the nanoparticles in the test media wasnot 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 andsubstrate-induced), microbial biomass C, enzymeactivities. Small shift in bacterial and protozoangene patterns by PCR-DGGE.

Tong et al., 2007

n-C60 No effect on respiration (basal), microbial biomassC (measured by substrate-induced respiration) andprotozoan abundance. Reduction in numbers ofbacteria. Small shift in bacterial and protozoangene patterns by PCR-DGGE.

Johansen et al.,2008

B) Carbon nanotubes

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

Lin and Xing,2007


root growth of ryegrass.

Metals

Aluminium No effect on seed germination of corn, cucumber,lettuce, radish, rape, and ryegrass. Rhizotoxic tocorn, lettuce, and ryegrass but stimulated radish andrape root growth.

Lin and Xing,2007

Zinc Reduced seed germination of ryegrass and reducedroot growth of corn, cucumber, lettuce, radish,rape, and ryegrass

Lin and Xing2007

Metal oxides

Al2O3 Phytotoxic (germination and seedling growth) butsee text.

Yang and Watts,2005

No effect on seed germination of corn, cucumber,lettuce, radish, rape, and ryegrass. No effect on rootgrowth of cucumber, lettuce, radish, rape, andryegrass. Reduced root growth of corn.

Lin and Xing,2007

TiO2 Stimulatory to spinach seed germination andseedling growth at low dose, phytotoxic at highdoses

Zheng et al., 2005

ZnO Reduced seed germination of corn and reduced rootgrowth of corn, cucumber, lettuce, radish, rape, andryegrass

Lin and Xing,2007

Yang and Watts (2005) reported the toxicity of alumina nanoparticles (13 nm, coatedwith and without phenanthrene) to root growth of five plant species (cabbage, carrot,corn, cucumber, and soybean) exposed to aqueous suspensions of the nanoparticles, butonly at high concentrations (2,000 mg/L). Loading of the alumina nanoparticles withphenanthrene reduced the toxicity of the nanoparticles. The nanoparticles were notphysically characterised prior to dosing, doses were not analytically confirmed, and in aletter to the Editor of Toxicology Letters, Murashov (2006) pointed out the experimentalprotocol of Yang and Watts (2005) did not distinguish toxicity caused by application ofthe aluminium in a nanoparticle form, and toxicity of solution aluminium derived fromthe nanoparticle. Indeed aluminium is a major component of soil minerals, known to bephytotoxic in acidic soils for almost a century (Magistad 1925) so the phytotoxicityobserved by Yang and Watts (2005) is not surprising, and clearly indicates the need toaccurately determine if the nanoparticulate form of a contaminant is toxic, or if thesoluble 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 seedgermination and early plant growth in simple Perlite media containing a completenutrient solution. Nano-TiO2 significantly increased seed germination and plant growthat low concentrations, but decreased these parameters at high concentrations. Bulk-


TiO2 had little effect. The manufactured nanoparticles in this study were not physicallycharacterised 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 plantspecies in simple aqueous media at pH 6.5–7.5. The nanoparticles were not physicallycharacterised prior to exposure and doses were not confirmed. The zinc-basednanoparticles 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 dependingon plant species. The authors attempted to quantify the solution zinc dose in theirexperiments by centrifugation (3000 G for 60 min) and filtration (0.7 μm). Theyreported that the centrifugation procedure did not fully separate the nanoparticles fromthe solution phase (assessed using TM-AFM), but they did not provide microscopicinformation on the solutions after filtration. Surprisingly, a 2000 mg/L suspension ofZnO after centrifugation and filtration returned a solution zinc concentration of only0.3-3.6 mg/L, significantly less than the concentration of Zn2+ in equilibrium with bulkZnO at pH 6.5–7.5, ~10–900 mg/L (Lindsay, 1979). Copper nanoparticles were alsorecently found to be potentially phytotoxic (Lee et al., 2008).

To date, there are only two reports in the literature of the terrestrial effects ofnanoparticles 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 ingranular form to soil microorganisms using soil respiration, microbial biomass,phospholipid fatty acid analysis, and enzyme activities as endpoints. The authors alsoexamined the DNA profile of the microbial community. All tests were performed in thelaboratory at optimal moisture conditions. In contrast to the observed microbial toxicityof n-C60 in vitro (Fortner et al., 2005), Tong et al. found no effect of n-C60 to anyendpoint in the soil medium used (silty clay loam, 4% organic matter, pH 6.9). Theysuggested 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 soilreducing the effective dose. A similar set of experiments was performed by Johansen etal. (2008), who examined the effect of n-C60 added to a neutral soil (pH 6.7) with loworganic C content (1.5%) on soil respiration, biomass C, bacterial and protozoanabundance and the PCR-DGGE profiling of bacterial and protozoan DNA. No effects ofexposure 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 smallchange in community structure, similar to the results of Tong et al. (2007). Similarresults from the same group were recently published for anaerobic bacteria typical ofwastewater sludge treatment systems (Nyberg et al. 2008).

There have been few reports of bioaccumulation or trophic transfer of nanomaterials tosoil invertebrates or mammals. A recent study of bioaccumulation of SWCNTs byearthworms indicated a very low bioaccumulation factor compared to pyrene (~100-foldlower) (Petersen et al., 2008), and a study of TiO2 accumulation by isopods (Porcellioscaber) also indicated a low bioaccumulation potential for these nanomaterials (Jemecet al., 2008).


These data highlight the need for more information on the interaction of nanoparticleswith soil components, and more quantitative assessments of aggregation/dispersion,adsorption/desorption, precipitation/dissolution, decomposition and mobility ofmanufactured nanoparticles in the soil environment. This information will aid theinterpretation of terrestrial ecotoxicity test data, and will inform the correct protocols forthe 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 mostnanomaterials 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 fortesting. It is therefore premature to suggest any regulatory limit for any nanomaterial insoils.

10. INTERNATIONAL PROGRESS ON NANOPARTICLE RISKASSESSMENT

10.1 International Approaches

With nanotechnology industries growing exponentially worldwide, the assessment ofthe 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 notnecessarily apply at the nanoscale, the current approach is still largely one ofinformation gathering through funding of additional research and development that willprovide a more sound basis than currently exists for managing the environmentalimpacts of manufactured nanomaterials.

The field is evolving extremely rapidly, and it is important to regularly check theliterature. CSIRO are part of an international Nanoparticles Advisory Group in theSociety of Environmental Toxicology and Chemistry that shares on a monthly basis thelatest research and regulatory developments, while the Nanosafety Theme in CSIRO’sNiche Manufacturing Flagship has close links with Dr Andrew Maynard of theWoodrow Wilson International Centre for Scholars (see below). Such linkages are vitalto both contributing to and accessing the latest information. CSIRO also has links intothe OECD Working Party on the Safety of Manufactured Nanomaterials, as discussedlater.

10.1.1 USA

In the US, a National Nanotechnology Initiative (NNI, 2001) was launched by theNational Science and Technology Council in 2001. Funding was provided to supportnanoscience 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. Environmentalissues were only of marginal concern. The National Science Foundation (NSF) later


established six facilities as part of Nanoscale Science and Engineering Centers. TheCenter for Biological and Environmental Nanotechnology at Rice University was thefacility 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 avehicle 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 withhow to apply the Toxic Substances Control Act to nanotechnology (Greenwood, 2007).A report prepared by the Woodrow Wilson Institute for Scholars investigated thedilemmas facing manufacturers and the USEPA in trying to deal with nanomaterialsunder that Act, using as an example, carbon nanotubes (WWIS, 2003). There weremany 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 madeunder the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA).

A nano risk framework was prepared in 2007 in a partnership between theEnvironmental Defense Fund and DuPont (Environmental Defense-DuPont, 2007). Theframework identified a basic set of environmental fate data including nanomaterialaggregation and disaggregation in the exposure media and screens for persistence andbiodegradability. For exposure assessments, they recommended acute toxicity andbioaccumulation testing, but identified a need for ecosystem level studies of effects onpopulations. Chronic tests would be required if a nanoparticle was potentially persistentand bioaccumulative. Depending on the fate, sediment testing might also be triggered.

The USEPA published a definitive Nanotechnology White Paper in 2007, following athree-year review, to inform EPA management of the science needs associated withnanotechnology. It included recommendations for addressing science issues andresearch needs (USEPA, 2007). More recently they produced a Draft NanomaterialResearch Strategy to guide the nanotechnology research program within the EPA’sOffice of Research and Development (USEPA, 2008). They identified four key researchthemes and seven key scientific questions which highlight the limitations of our currentknowledge:

1. Sources, Fate, Transport and Exposure

a. Which nanomaterials have a high potential for release from a life-cycleperspective?

b. What technologies exist, can be modified, or must be developed to detectand quantify engineered nanomaterials in environmental media andbiological samples?


c. What are the major processes/properties that govern the environmentalfate of engineered nanomaterials, and how are these related to physicaland chemical properties of these materials?

d. What are the exposures that will result from the releases of engineerednanomaterials?

2. Human Health and Ecological Research to Inform Risk Assessment and TestMethods

a. What are the effects of engineered nanomaterials and their applicationson human and ecological receptors, and how can these effects be bestquantified and predicted?

3. Risk Assessment Methods and Case Studies

a. Do Agency risk assessment approaches need to be amended toincorporate special characteristics of engineered nanomaterials?

4. Preventing and Mitigating Risks

a. What technologies or practices can be applied to minimize risks ofengineered nanomaterials through their life cycle, and how cannaonotechnology’s beneficial uses be maximised to protect theenvironment?

10.1.2 United Kingdom

The Royal Society and Royal Academy of Engineering released a report in 2004 onnanoscience and nanotechnologies that addressed the current state of environmentalassessment of nanomaterials. It proposed that nanoparticulate forms of chemicalsshould be treated as new chemicals for regulatory purposes, and identified the need fornew research to determine routes of exposure and toxicity. The UK government hasreleased several reports investigating the potential risks posed by manufacturednanoparticles (DEFRA, 2005, 2007). The reports place the UK research programoverseen by a cross-government Nanotechnology Research Coordination Group in aninternational context. They are collaborating with the OECD and the InternationalStandards Organisation (ISO) to share data and experiences to maximise the speed withwhich potential risks can be identified and managed.

Specific task forces are addressing: (i) Metrology, characterisation, standardisation andreference materials, (ii) Exposures: sources, pathways and technologies, (iii) Humanhealth 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 ofTrade and Industry identified a number of gaps in the application of environmentalregulations to nanomaterials. A lack of knowledge of toxicity data was a critical issue.


10.1.3 Other International Activities

The OECD’s Environment Directorate has been active in sponsoring a number ofmeetings dealing with the safety of manufactured nanomaterials. Details of these areavailable on their website (http://www.oecd.org/ehs). Reports on developments inChina, 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 thefollowing 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 MaxineMcCall and Simon Apte of CSIRO and NICNAS staff, a sponsorship program wasinitiated whereby member countries volunteered to undertake work on specificnanomaterials of national interest in collaboration with each other. Australia agreed toundertake the study of zinc oxide, cerium dioxide and silver.

Nanoparticles are fully covered by REACH, the new European Community regulationon chemicals and their safe use requirements. One of the first activities of the memberState Committee under the European Chemicals Agency (ECHA) was to institute aNanoparticles Working Group. Nominations for this Working Group were sent toECHA by member states and observers from industry and other countries including theUS. .

The European Chemical Industries Council (CEFIC) is currently reviewing strength andweaknesses of the REACH risk assessment framework for nanoparticles, building onthe EU SCENIHR (Scientific Committee on Emerging and Newly Identified HealthRisks) 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 datacollection and the need to benefit from European and US experiences.

http://www.oecd.org/ehs


10.2 Australian Activities

Australia has been following a path similar to other major international players in itsapproach to nanomaterial risk assessment. NICNAS, the national regulator of industrialchemicals, issued a voluntary call for information to importers and manufacturers ofnanomaterials in 2006, as discussed earlier. Industry was asked to provide informationon uses and quantities of nanomaterials imported or manufactured for industrialpurposes, including use in cosmetics and personal care products. The information wasdesigned to assist in understanding which nanomaterials are available on the market orclose to commercialisation, and help focus our efforts to ensure the adequacy of theregulatory scheme to assess nanomaterials.

Nanomaterials used exclusively as therapeutic goods, pesticides or food additives do notfall within the scope of NICNAS, and were consequently outside the voluntary call forinformation. The results of its findings were published on its websitehttp://www.nicnas.gov.au. Information from a new call is currently being compiled.

Chemicals that are not listed on the Australian Inventory of Chemical Substances(AICS), which is based on the chemical formula and CAS number of chemicals (with nosize definition), are generally regarded as "new" and must be notified to NICNAS andassessed 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 chemicalformula and CAS number) are currently considered to be "existing" chemicals. Thesenanoscale existing chemicals can be selected for assessment if they potentially present achanged risk of adverse health and/or environment effects. To date, NICNAS has notassessed any nanomaterials with novel properties.

NICNAS is currently examining the suitability of its regulatory framework andprocesses to protect human health and the environment in association with the OECDWPMN, and by engagement with Australian government agencies under the NationalNanotechnology Strategy. At the same time, NICNAS has convened a NanotechnologyAdvisory Group which has three members each from the community and industry, andtwo members from academia and one from NICNAS, and which NICNAS chairs.

A national Nanotechnology Roundtable was hosted by the National Health and MedicalResearch Council (NHMRC) in December 2006. The Roundtable was attended byrepresentatives from Australian academic institutions, health and environmentgovernment departments, regulatory bodies and industry and a representative from theNew Zealand Health Research Council. The Chief Executive Officer of the NHMRC isusing the outcomes of the Roundtable to inform future directions for the NHMRC.

In June 2006, the National Nanotechnology Strategy Taskforce produced a report for thegovernment on "Options for a National Nanotechnology Strategy" (NNST, 2006). Itsmajor findings with respect to the environment were:

nanomaterials do exhibit novel properties that will have health safety andenvironmental implications, but the significance is unclear at the moment, andcannot be easily predicted due to a gap in knowledge;

http://www.nicnas.gov.au


measurement and assessment of nanoparticles is a key priority for furtherresearch;

there is a need for continued international cooperation in the field;

the Australian regulatory system needs to be flexible to address the challengesposed by nanoparticles, and that coordinated processes at the regulatory levelneed to be put in place; and

while no serious risk is evident now, potential real risks in the use ofnanotechnologies in Australia must be identified so that appropriate riskmanagement strategies can be employed for their safe use.

Following on from the Taskforce’s report, a HSE Working Group comprising federalagencies with responsibility for policy and implementation of Australia's regulatoryframeworks was established to consider HSE issues in more detail. This groupcommissioned a review of the capacity of Australia's regulatory frameworks to manageany 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 theuse of nanoparticulate zinc oxide and titanium dioxide in sunscreens, concluding thatthey did not represent a major health threat. At that time, Food Standards AustraliaNew Zealand (FSANZ) had not received any applications to consider the regulation ofany nanomaterials under the Australia New Zealand Food Standards Code.

The APVMA have also recently published in the Gazette, a voluntary call forinformation on nanomaterials in agricultural or veterinary chemicals, or agricultural andveterinary chemical products. APVMA has published a position paper onnanotechnology (APVMA, 2008).

The Department of Infrastructure, Transport, Regional and Local Government (transportof hazardous materials) and the Department of the Environment, Water, Heritage andthe Arts (DEWHA) in conjunction with the above bodies are currently assessinginternational research in this area.

11. DEVELOPMENT OF TECHNICAL GUIDELINES FORNANOMATERIAL ASSESSMENT

An Environmental Risk Assessment Guidance Manual for industrial chemicals wasissued by the then Department of Environment and Water Resources (now DEWHA) in2007 (DEW, 2007). While this did not consider nanomaterials, the draft frameworkoutlined an approach that was consistent with NChEM, the discussion paper on anational framework for chemicals management in Australia prepared by theEnvironment Protection and Heritage Council (EPHC, 2006).


11.1 Exposure Assessment Incorporating Nanomaterial Fate

The development of appropriate technical guidelines that cover the impacts ofmanufactured nanomaterials in waters, sediments and soils first requires an evaluationof the fate of nanomaterials in these environments. Only then can exposure parametersbe appropriately assessed for comparison against any available environmental qualityguidelines. It is evident from the information presented in this report, that the fate ofnanomaterials in the environment, and their toxicity to biota, is likely to be a function ofsize, shape, surface properties and bulk composition. Inferring fate and toxicity frombulk composition alone (e.g. zinc oxide, CNTs, nanodots) is likely to be inappropriatedue to the wide range of surface functional coatings applied to many nanomaterials.Surface properties will therefore play a key role in interactions with all environmentalmatrices (soils, sediments and waters).

It is possible to lay out the following key questions that incorporate basic considerationsof 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 aquaticsystems, noting that the behaviour may differ for different product formulations:

(i) What is the particle size of dispersed nanomaterials in a natural receivingwater system (i.e. extent of aggregation) as determined by appropriatetechniques? The key factor here is the existence of dispersed or aggregatedparticles that are in the nano range (<100 nm) and likely to have propertiesdiffering 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 bydialysis or ultrafiltration) and/or dissociated thereby having potentiallydifferent biological availability to the insoluble fraction?

(iv) Does the soluble fraction change with time, and if so over what timescale?


(v) What is the estimated mass concentration of the insoluble aggregateddispersion in the nano size range? If the concentration is low, areinteractions with natural colloids at higher concentrations likely to modifythe form of the nanomaterial through adsorption.

C. Fate in Sediments

If nanomaterials accumulate in sediments, it would be assumed that this is theconsequence of settling from the water column due to a loss of buoyancy. To do sorequires considerable aggregation of particles or association with other water-borneparticulates notionally to exceed 1 µm in size, although this is density dependent. Onceaccumulated in the sediments, the key fate questions relate to dissolution andtransformation. The key questions might be:

(i) Is there any dissolution of nanomaterials in the sediment pore waters underthe 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 thereany nano-sized manufactured particles that might pose a different threat tonatural nanoparticles, or natural or manufactured macroparticles.

D. Fate in Soils

As already noted, nanomaterial interactions in soils are poorly characterised. In terms ofkey 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 inthat form? How is this solubility affected by soil pH, salinity, sodicity, redoxconditions and time?

(ii) What is the particle size of nanomaterials in soil after any naturalaggregation processes? How easily are nanomaterials sorbed and retained bysoil minerals and organic matter?

(iii) Are nanomaterials more mobile through soils than natural nanoparticles?Can they be vectors for enhanced transport of contaminant solutes togroundwaters, 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 thesecond component of any hazard assessment, the effects assessment. It will dictate theevaluation of potential toxicity based on the existing toxicity database and a knowledgeof how changes to basic nanomaterials in the environment affect their bioavailability.


In terms of defining guideline concentrations for nanomaterials in waters, sediments andsoils, data are required from an appropriate set of toxicity tests (DEW, 2007) that can beused to derive an appropriate PNEC. For aquatic biota, data are needed from at leastfive species from four taxonomic groups if a statistical extrapolation method is to beapplied (ANZECC/ARMCANZ, 2000), although this size dataset must be viewed as aminimum. Alternatively a low reliability value can be estimated by the application of anassessment factor to the lowest chronic no observed effects concentration (NOEC). Forwater sample toxicity data, as shown in Table 10, this will currently be the onlyalternative.

The limitations of these toxicity datasets have been discussed earlier, the principalconcern being that the nanomaterials used are appropriately characterised in terms ofkey parameters, especially particle size (degree of aggregation), and any specificformulations 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-specificguidelines that take into account environmental chemistry and its effect on nanoparticlebehaviour (especially the effect of ionic strength on aggregation).

Based on findings to date, the following hypotheses are suggested with respect to thebioavailability and potential toxicity of nanomaterials in the environment:

(i) The fate and bioavailability of a particular nanomaterial might be expected tochange in the presence of additives that affect surface properties;

(ii) Small dispersed nanoparticles (<20 nm) are likely to be more bioavailableand potentially toxic than large aggregates (>100 nm). Aggregates cannevertheless exert toxicity compared with bulk material especially ininstances related to ROS generation where the aggregate surface area may beonly marginally lower than that of its composite nanoparticles.

(iii) Interaction with other particles and aggregation will be dependent on boththe surface charge and the surface area of nanoparticles. Low surface areaand net negatively charged particles are less prone to aggregation and arepotentially more mobile (but perhaps less bioavailable) since mostmembranes have negative trans-membrane potentials;

(iv) Interaction of nanoparticles with other naturally present colloids or organicmacromolecules will also affect reactivity. Such interactions will befavoured by high nanoparticle surface areas and excess concentrations ofnatural colloids, and would be expected to result in larger particles withreduced bioavailability;

(v) Aggregation is faster in environments of high ionic strength, i.e. highhardness or saline waters, saline soils, or saline sediments;

(vi) With greater aggregation, particle toxicity approaches that of the equivalentbulk macroparticles;


(vii) For metallic nanomaterials, water-soluble and potentially dissociatednanoparticles are likely to be more toxic than insoluble and undissociatedparticles of the same material, although there are mechanisms by whichinsoluble materials can exert toxicity; and

(viii) The kinetics of both aggregation and dissolution will influence the abovetoxicity.

Current indications are that it will only be possible to provide low reliability PNECs fora limited set of nanomaterials, however, an awareness of the factors affecting fate willguide the design of field-relevant toxicity testing.

11.3 Possible Approaches to Environmental Hazard Rankingof Nanomaterials

As has been discussed, there are currently too few data available to set appropriatelimits or environmental guideline threshold numbers that accommodate all aspects ofthe various formulations, and their fate in the environment, although an approach hasbeen outlined that might allow these to be achieved. It might then be appropriate todevelop a basic matrix approach to assess and rank the potential environmental mobilityand hazard from nanomaterials, where bulk size, solubility and surface properties areconsidered. As yet, there are insufficient data to develop such an approach.

The prediction of behaviour based on the physical or chemical properties ofnanomaterials is only possible in a very limited way. For example, the use ofquantitative structure activity relationships (QSARs) may be useful for a limited numberof comparable nanomaterial types whose structures differ in the nature of chemicalsubstitution on a base molecular structure (e.g. fullerenes or CNTs). Nothing has yetbeen published in this area, although the Joint Research Centre (JRC) of the EuropeanCommission in late 2006 funded the Computational Toxicology Group, within theEuropean Chemicals Bureau in Ispra, Italy, to review the applicability of (Q)SARs tonanoparticles (http://ecb/jrc.it/QSAR/). While no publications have yet appeared, theJRC website indicates that their activities are focused on the development andharmonisation of methods for toxicity testing of nanomaterials, the in vitro test of arepresentative set of manufactured nanomaterials on critical cell lines and encompassrelated studies on nanometrology and reference materials as well as the development ofdatabases and studies on the applicability of 'in silico' methods adapting the traditionalQSAR paradigm.

A JRC report (Dearden and Worth, 2007) outlines the concept of QSPRs (quantitativestructure property relationships) as a cost-effective computational alternative tomeasurement of fate and toxicity. This is being explored in relation to nanomaterials. Itis important that these models go as far as predicting actual fate rather than stoppingonly with the raw material and not its in-field characteristics.

http://ecb/jrc.it/


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 ofthe stability, and short- and long-term fate of the various likely formulations thatmight reach these compartments of the environment, and the development oftechniques to distinguish natural from manufactured nanoparticles. Thesemeasurements should focus on particle concentration, size and surfacecharacteristics (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 levelsas required to derive PNECs using species sensitivity distributions. It is criticalthat appropriate verification of particle and solute dose be undertaken in allecotoxicity testing, necessitating significant effort in (1) above.

3. As a precursor to toxicity testing, it will be necessary to develop standard (andvalid) methodologies for the hazard ranking of nanomaterial toxicity. These willneed to ensure the stability of the nanoparticle suspensions over the duration ofthe standardised toxicity tests.

4. Comparisons of toxicity testing in natural vs. synthetic water and soil samplesdemonstrating the effects of natural colloids.

13. ACKNOWLEDGEMENTS

The authors acknowledge Drs Natasha Franklin, Nicola Rogers and Simon Apte foruseful discussions and information provided for this report. We are grateful to Dr GlenWalker (DEWHA) for his careful and comprehensive refereeing. The project wascommissioned 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..

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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 ananoscale, multistep fabrication process. Each step results in a new “generation” that has twicethe 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 tobring 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


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 thathave been created due to deliberate human activity.

Microemulsion: An emulsion (dispersion of one immiscible liquid in another) where theparticles 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 theparticles in the dispersed phase are less than 1000 nm. Nanoemulsions are kinetically but notthermodynamically stable.

Nanomaterials: Materials having structured components that have one dimension lower than100 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 intonanoparticles, 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 nmto 100 nm play a critical role.

Nanotube: A one-dimensional fullerene (a convex cage of atoms with only hexagonal and/orpentagonal faces) with a cylindrical shape.

Nanowires: One-dimensional structures, with unique electrical and optical properties, that areused 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, orchloroplast, 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. Thequantum dot is considered to have greater flexibility than other fluorescent materials, whichmakes it suited to use in building nano-scale computing applications where light is used toprocess information. They are made from a variety of different compounds, such as cadmiumselenide.


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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Fate of Manufactured Nanomaterials in the Australian Environmentenvironment.gov.au/system/files/pages/371475a0-2195-496d... · Fate of Manufactured Nanomaterials in the Australian