ecotoxicity of engineered nanoparticles to freshwater

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Ecotoxicity of engineered nanoparticles to freshwater organisms

Hartmann, Nanna Isabella Bloch

Publication date:2011

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Citation (APA):Hartmann, N. I. B. (2011). Ecotoxicity of engineered nanoparticles to freshwater organisms. Technical Universityof Denmark.

https://orbit.dtu.dk/en/publications/c0a366db-32b7-4483-8b95-0ab9828d2c70

PhD ThesisApril 2011

Nanna Bloch Hartmann

Ecotoxicity of engineered nanoparticles to

freshwater organisms



PhD Thesis

April 2011

DTU Environment

Department of Environmental Engineering

Technical University of Denmark

DTU Environment



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April 2011

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PhD Thesis,

The thesis will be available as a pdf-file for downloading from the homepage of

the department: www.env.dtu.dk

April 2011

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Preface This PhD thesis presents research into the environmental effects of engineered nanoparticles undertaken between September 2007 and January 2011 at the Department of Environmental Engineering, Technical University of Denmark (DTU) under supervision of Associate Professor Anders Baun. The project was co-supervised by Senior Scientist Mona-Lise Binderup at the National Food Institute, DTU. The thesis consists of five chapters, covering three main topics: I) The current level of scientific knowledge regarding ecotoxicological effects of engineered nanoparticles, II) Exploratory studies into the potential of nanoparticles as carriers for co-existing pollutants and III) Examination of the applicability of commonly used ecotoxicity test methods to nanoparticles. The thesis is based on seven scientific papers, which cover the major findings of the PhD project:

I. Hartmann, N.B., von der Kammer, F., Hofmann, T., Baalousha, M., Ottofuelling, S., & Baun, A., 2010. Algal testing of titanium dioxide nanoparticles – Testing considerations, inhibitory effects and modification of cadmium bioavailability. Toxicology, 269 (2-3) 190–197.

II. Hartmann, N.B. & Baun, A., 2010. The nano cocktail: ecotoxiological effects of engineered nanoparticles in chemical mixtures. Integrated Environmental Assessment and Management, 6 (2) 311–313.

III. Hartmann, N.B., Buendia, I., & Baun, A., 2011. Degradability of aged aqueous suspensions of C60 nanoparticles. Submitted to Environmental Pollution. Under revision.

IV. Hartmann, N.B., Engelbrekt, C., Zhang, J., Ulstrup, J., Kusk, K.O., & Baun, A., 2011. The challenges of testing insoluble metal and metal oxide nanoparticles in algal bioassays: titanium dioxide and gold nanoparticles as case studies, Manuscript

V. Hartmann, N.B., Legros, S., von der Kammer, F., Hofmann, T., & Baun, A., 2011. The potential of TiO2 nanoparticles as carriers for heavy metal uptake in Lumbriculus variegatus and Daphnia magna. Manuscript

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VI. Baun, A., Hartmann, N.B., Grieger, K.D., & Kusk, K.O., 2008. Ecotoxicity of engineered nanoparticles to aquatic invertebrates: a brief review and recommendations for future toxicity testing. Ecotoxicology, 17 (5) 387-395.

VII. Baun, A., Sørensen, S.N., Rasmussen, R.F., Hartmann, N.B., & Koch, C.B., 2008, Toxicity and bioaccumulation of xenobiotic organic compounds in the presence of aqueous suspensions of aggregates of nano-C60. Aquatic Toxicology 86 (3) 379–387.

In addition, the following publications, co-authored in 2007-2011 and related to the topic of this PhD project, are not included in this thesis: Stone, V., Hankin, S., Aitken, R., Aschberger, K., Baun, Anders, Christensen, F., Fernandes, T., Hansen, S.F., Hartmann, N.B., Hutchinson, G., Johnston, H., Micheletti, C., Peters, S., Ross, B., Sokull-Kluettgen, B., Stark, D., Tran, L., 2010. Engineered nanoparticles: Review of health and environmental safety (ENRHES). Project Final Report, European Commission, FP7 CSA #21843. Grieger, K.D., Fjordbøge, A., Hartmann, N.B., Eriksson, E., Bjerg, P.L., Baun, A., 2010. Environmental benefits and risks of zero-valent iron nanoparticles (nZVI) for in situ remediation: Risk mitigation or trade-off? Journal of Contaminant Hydrology, 118, 165-183. Baun, A., Hartmann, N.B., Grieger, K.D., Hansen., S.F., 2009. Setting the limits for engineered nanoparticles in European surface waters – are current approaches appropriate? Journal of Environmental Monitoring, 11, 1774-1781. Navarro, E. Baun, A., Behra, R., Hartmann, N.B., Filser, J., Miao, A-J., Quigg, A., Santschi, P.H., Sigg, L., 2008. Environmental behavior and ecotoxicity of engineered nanoparticles to algae, plants, and fungi. Ecotoxicology, 17, 372-386. The papers above are not included in this www-version but can be obtained from the library at DTU Environment. Contact info: Library, Department of Environmental Engineering, Technical University of Denmark, Miljoevej, Building 113, DK-2800 Kgs. Lyngby, Denmark or [email protected].

February 2011 Nanna Bloch Hartmann

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Acknowledgements

I would like to thank everyone who has, in one way or another, contributed to the completion of this PhD project. In particular I wish to express my appreciation to the people who have provided technical, experimental and graphical assistance as well as to collaborators and co-authors. A special thanks to my supervisor, Anders Baun, for inspirational discussions, enthusiastic guidance and mentorship. Thank you to all my present and former colleagues at DTU Environment for providing such a positive and inspiring working atmosphere. In particular I would like to thank my colleagues in the research group on Nanotechnology & Risk as well as the technical staff. Heartfelt gratitude goes to my friends for their encouragement and patience. Finally I would like to thank my family for their loving support and for being my foundation. “There is no such thing as a failed experiment, only experiments with unexpected outcomes” (Richard Buckminster Fuller)

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Summary A large variety of societal benefits are expected from the development and use of engineered nanoparticles. At present, the majority of ‘nano-products’ put on the market can be classified as consumer products, whereas future applications are expected to have more widespread and societal benefits in areas as diverse as cancer treatment, groundwater remediation and industrial coatings. Nanoparticles are used to give the products new and improved characteristics. Yet exactly these new and nano-specific properties might be a cause of concern in a health and environment context. In order to ensure adequate protection of humans and the environment, a pro-active effort to understand, identify and minimise potential risks is needed at an early stage in the innovation process. However, due to the fundamentally different nature of nanoparticles as discrete entities, compared to ‘conventional’ water soluble chemicals, many aspects of commonly used test methods for evaluation of potential adverse environmental effects make their applicability to nanoparticles questionable. For this reason the overarching aim of this PhD project has been to acquire information, which can be disseminated and applied in relation to appropriate test methods for identifying potential adverse effects of nanoparticles; this is of great relevance from both a scientific and regulatory point of view. An important aspect of this project was the acquisition of experience in testing nanoparticles in aqueous test systems – both through practical lab-based studies as well as literature studies. The process of testing nanoparticles in aquatic ecotoxicity tests has been as important in the project approach as the test outcomes. Applied test methods have included acute and chronic toxicity tests as well as bioaccumulation studies with freshwater filter feeder Daphnia magna, sediment feeder Lumbriculus variegatus and green alga Pseudokirchneriella subcapitata. The results made it possible to identify major scientific and methodological challenges in the testing of nanoparticles compared to ‘conventional’ chemicals. It has been highlighted that while it is possible to obtain dose-response relationships for nanoparticles, such tests often raise as many questions as they answer. Issues requiring further attention include non-linear concentration-aggregation relationships, adhesion of nanoparticles to the cell and organism surfaces, problems of relating effects directly to properties of the primary nanoparticles, dynamic two-way interactions between nanoparticles and organisms as well as how to best quantify and qualify exposure in a dynamic system. Particularly for algal growth inhibition tests, methodological issues were raised related to biomass quantification methods. Here, a combination of several

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techniques is recommended. The combination of visual inspection, cell counting and fluorescence measurement of acetone-extracted pigments was found to give additional insight into the nature of the observed effects. A separate task was to explore the role of nanoparticles in chemical mixtures, elucidated through experimental and conceptual studies. Interaction between nanoparticles and co-contaminants in chemical mixtures may result in changes in the bioavailability and toxicity of the individual compounds. A conceptual model for interaction scenarios between nanoparticles and co-existing environmental pollutants was developed. Experimental results showed that several types of nanoparticles (e.g. TiO2 and C60) have a large adsorption capacity for some heavy metals and organic chemicals. Nonetheless, cadmium adsorbed to TiO2 nanoparticles was found to be bioavailable to D. magna, L. variegatus and P. subcapitata. TiO2 nanoparticles were seen to attach to the surface of P. subcapitata and to mainly be located in the gut of D. magna. In D. magna the presence of TiO2 resulted in increased uptake of cadmium but not in overall increased toxicity. Interaction studies (especially binary multiple-dose studies) may be considered premature at present due to the additionally increased test system complexity. This may hamper the interpretation of test results both compared to interaction studies for conventional water-soluble chemicals and ecotoxicological tests of only nanoparticles. However, interaction studies using single doses of nanoparticles may still provide indications of the potential of nanoparticles to influence mixture toxicity and how this occurs. Also, when fundamental general test procedures for testing of nanoparticles are in place, the role of nanoparticles in chemical mixtures is an important field for future studies. In the light of these findings, it is recommended that research in the field of nanoecotoxicology is prioritised towards the methodological challenges that have to be overcome in order to obtain meaningful results. Compared to conventional water-soluble chemicals, additional considerations regarding test procedures are needed to gain an insight into the underlying mechanisms. The influence of particle behaviour in the test system on observed effects and actual toxic mechanisms remains to be explored further. As a supplement to traditional endpoints such as mortality, reproduction and bioaccumulation, it is crucial to perform exploratory in-depth investigations of biological processes and analytical methods to ensure the robustness and applicability of test methods and test endpoints. A focus on methodological problems, test system dynamics, ecotoxicokinetics and effects mechanisms should therefore be encouraged.

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Dansk sammenfatning Udvikling og anvendelse af industrielt fremstillede nanopartikler forventes at resultere i en lang række samfundsmæssige fordele. På nuværende tidspunkt kan de fleste ’nano-produkter’ på markedet betegnes som forbrugsvarer, mens fremtidige anvendelser forventes at have bredere samfundsmæssige fordele indenfor områder så forskellige som kræftbehandling, vandrensning og industrielle belægninger. Nanopartikler anvendes til at give produkterne nye og forbedrede karakteristika. Netop disse nye og nano-specifikke egenskaber kan dog samtidig give anledning til sundheds- og miljømæssige overvejelser. For at sikre tilstrækkelig beskyttelse af mennesker og miljø er en pro-aktiv indsats for at forstå, identificere og minimere potentielle risici nødvendig på et tidligt tidspunkt i innovationsprocessen. Nanopartikler er fundamentalt forskellige fra ‘konventionelle’ vandopløselige kemikalier i kraft af deres natur som diskrete enheder. Derfor er der flere aspekter af gængse testmetoder til vurdering af potentielle negative miljømæssige effekter, som gør deres anvendelighed for nanopartikler tvivlsom. Det overordnede mål med dette PhD projekt har således været at tilvejebringe informationer, som kan formidles og anvendes i forhold til velegnede testmetoder til identifikation af potentielle negative effekter af nanopartikler. Dette er af stor betydning både ud fra et videnskabeligt og et regulatorisk synspunkt . Et vigtigt aspekt af dette projekt var at opnå erfaringer i forbindelse med testning af nanopartikler i akvatiske testsystemer – både gennem laboratorie-baserede of litteratur-baserede studier. Selve processen at teste nanopartikler i akvatiske økotoksikologiske tests har været ligeså væsentligt i tilgangen til projektet som testresultaterne i sig selv. De anvendte testsmetoder har inkluderet tests for akut og kronisk toksicitet samt bioakkumuleringsstudier med ferskvandskrebsdyr Daphnia magna, sedimentorm Lumbriculus variegatus og grønalge Pseudokirchneriella subcapitata. Resultaterne har gjort det muligt at identificere de vigtigste videnskabelige og metodiske udfordringer ved test af økotoksicitet af nanopartikler i forhold til ‘konventionelle’ kemikalier. Det er blevet fremhævet, at mens det er muligt at etablere dosis-respons sammenhænge for nanopartikler, så rejser sådanne forsøg ofte ligeså mange spørgsmål som de besvarer. Spørgsmål, der kræver yderligere opmærksomhed, omfatter ikke-lineære sammenhænge mellem koncentration og aggregering, vedhæftning af nanopartikler til celle- og organismeoverflader, problemer ved at relatere effekter direkte til egenskaber ved de primære nanopartikler, dynamiske to-vejs interaktioner mellem nanopartikler og organismer samt hvordan man bedst

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kvantitativt og kvalitativt beskriver eksponering i et dynamisk system. Især ved test af algevækst-inhibering påpeges metodiske spørgsmål i forbindelse med metoder til kvantificering af biomasse. Her anbefales en kombination af flere teknikker. Kombinationen af visuelle undersøgelser, celletælling og fluorescensmåling af acetone-ekstraherede pigmenter viste sig at give yderligere indblik i karakteren af de observerede virkninger. En separat del af projektet var undersøgelse af nanopartiklers rolle i kemiske blandinger, belyst gennem eksperimentelle og konceptuelle studier. Interaktioner mellem nanopartikler og andre miljøfremmede stoffer i kemiske blandinger kan resultere i ændringer i biotilgængeligheden og toksiciteten af de enkelte stoffer. En konceptuel model for interaktions-scenarier mellem nanopartikler og kemikalier blev udviklet. Forsøgsresultater viste, at flere typer af nanopartikler (f.eks TiO2 og C60) har en stor adsorptionskapacitet for visse tungmetaller og organiske kemikalier. Ikke desto mindre blev cadmium adsorberet til TiO2 nanopartikler fundet at være biotilgængeligt for D. magna, L. variegatus og P. subcapitata. TiO2 nanopartikler sås at vedhæfte til overfladen af P. subcapitata og primært være lokaliseret i fordøjelseskanalen i D. magna. I D. magna resulterede tilstedeværelsen af TiO2 i en øget optagelse af cadmium, men øgede ikke den samlede toksicitet. Undersøgelser af kombinationseffekter (især for binære blandinger med flere doser) kan synes forhastet på nuværende tidspunkt, da sådanne studier medfører yderligere forøget kompleksitet af testsystemet. Dette kan besværliggøre fortolkningen af testresultater – både set i forhold til undersøgelse af kombinationseffekter for konventionelle vandopløselige kemikalier og i forhold til økotoksikologiske tests af nanopartikler alene. Ved at foretage forsøg med en fastholdt koncentration af nanopartikler kan sådanne forsøg dog stadig give indikationer af de potentielle effekter af nanopartikler i kemiske blandinger og om årsagen til sådanne effekter. Når grundlæggende testprocedurer til testning af nanopartiklers effekter er på plads, kan nanopartiklers rolle i kemiske blandinger være et vigtig område for fremtidige undersøgelser. I lyset af disse resultater anbefales det, at forskning indenfor nano-økotoksikologi prioriteres i retning af de metodiske udfordringer, som skal løses for at kunne opnå meningsfulde resultater. Set i forhold til ‘konventionelle’ vandopløselige kemikalier er yderligere overvejelser omkring testprocedurer nødvendige for at få et indblik i underliggende effekt-mekanismer. Betydningen af partiklernes opførsel i testsystemet i forhold til de observerede effekter og effekt-mekanismer skal fortsat undersøges nærmere. Som et supplement til traditionelle biomarkører

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såsom dødelighed, reproduktion og bioakkumulering, er det afgørende at udføre eksplorative, dybdegående undersøgelser af biologiske processer og analysemetoder for at sikre soliditet og anvendelighed af testmetoder og biomarkører. Fokus på metodiske problemer, dynamikken i testsystemet, ecotoxico-kinetik og virkningsmekanismer bør derfor fremmes.

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Table of Contents Preface .................................................................................................................... i Acknowledgements ..............................................................................................iii Summary ............................................................................................................... v

Dansk sammenfatning ........................................................................................ vii 1. Background and aims ....................................................................................... 1

2. Nanoparticles and environmental concerns ................................................... 3

2.1. What are nanoparticles? .............................................................................. 3

2.2. Applications ................................................................................................. 3

2.3. Environmental exposure .............................................................................. 6

2.4 Currently available guidance for ecotoxicity testing of nanoparticles ......... 8

3. Ecotoxicity, uptake and behaviour of metal, metal oxide and carbon-based nanoparticles in aquatic test systems ................................................................ 11

3.1. Direct and indirect effects of metal oxide nanoparticles in algal growth inhibition tests .................................................................................................. 12

3.2. Intracellular uptake in algae ...................................................................... 16

3.3. Particle versus metal ion effects ................................................................ 17

3.4. Effects resulting from impurities and solvents .......................................... 19

3.5. Uptake of nanoparticles into the gut of filter and sediment feeders .......... 20

3.6. Biomodification of nanoparticle characteristics and behaviour ................ 26

4. Nanoparticles in chemical mixtures and the consequences for bioaccumulation and toxicity ............................................................................. 31

5. Test methods for hazard identification –current applicability and future challenges ............................................................................................................. 39

5.1. Behaviour of nanoparticles in aquatic test systems ................................... 39

5.2. Exposure quantification in a dynamic system ........................................... 44

5.3. Correlating particle characteristics to effects ............................................ 46

5.4. Adaptation of current testing methods – what is the potential? ................ 49

6. Conclusion ....................................................................................................... 55

6.1. Outlook ...................................................................................................... 55

References ............................................................................................................ 57

Appendices .......................................................................................................... 71

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1. Background and aims The use of engineered nanoparticles in a large variety of applications is expected to result in a wide range of societal benefits ranging from nanomedicines to electronics and environmental remediation – all of which are enabled by the ability to manipulate materials at nano scale. Nanoparticles can be produced with very specific characteristics targeted precisely at their specific applications. This may be achieved through specific particle sizes, chemical compositions, surface coatings or functionalisations. Yet exactly the properties of nanoparticles, which enable their use in these novel applications, are also those which might be a cause for concern in a health and environment context such as persistency, biocompatibility, high chemical reactivity and antimicrobial properties. With increasing industrial production and application of nanomaterials, the issue of innovation versus environmental and human health risks becomes increasingly important. Despite increasing focus on potential environmental hazards of engineered nanoparticles, progress in understanding potential environmental impacts of nanoparticles is lagging far behind the development of new particle types (Bernhardt et al., 2010). The first aim of this thesis is therefore to review current knowledge regarding the ecotoxicity and bioaccumulation of engineered nanoparticles with focus on particles composed of metal oxides (in particular TiO2), metals (in particular gold) and carbon (in particular C60) (addressed in Chapter 3 and Papers III & VI). The review is limited to aquatic freshwater ecotoxicity with emphasis on algae, crustaceans and sediment-dwelling worms, reflecting the overall topic of this thesis and the work undertaken during the PhD project period. At the same time these species are significant both for the ecological food web and as key organisms in regulatory testing (Baun et al., 2008a – Paper VI). Through discussions of specific issues of interest, this review aims to describe our current state of knowledge as well as highlight potential relationships between particle properties and observed effects, while also drawing attention to methodological difficulties, knowledge gaps and uncertainties. As is the case for conventional chemicals, engineered nanoparticles exist in the aquatic environment as part of a complex mixture. Mixture interactions will take place in particle production, product manufacturing and use and after disposal. These interactions are likely to influence bioavailability of the individual compounds and hence their ecotoxicological effects. Interaction scenarios for

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nanoparticles and co-existing chemicals differ from traditional mixture toxicity scenarios due to the presence of a solid phase. Furthermore, nanoparticle properties may be favourable for interactions through sorption due to a large surface area per mass and a small size, which may allow uptake into organisms, organs and cells. The second aim of this thesis is therefore to evaluate the role of nanoparticle interactions with co-existing pollutants with particular focus on their role as pollutant carriers (addressed in Chapter 4 and Papers I, II, V & VII). A fundamental pre-requisite for understanding and discussing the potential negative environmental effects of engineered nanoparticles is the ability meaningfully to conduct and interpret tests for their hazard potential. The current test paradigm for determining adverse ecotoxic effects is based on the assumption that the chemicals are water soluble, which will not generally be the case for nanoparticles. The third aim of this thesis is therefore to examine the applicability of commonly used ecotoxicity test methods to nanoparticles (addressed in Chapter 5 and Papers I & IV). The purpose of this section is not to suggest specific changes to standard test methods for regulatory testing but is rather a scientific discussion, which may at a later point feed into amendments and development of nano-specific test guidelines. This section will focus primarily on algal growth inhibition test methods but will also cover general problems related to exposure, dose-metrics, test artefacts etc. in freshwater toxicity testing. Hence in Chapter 5 essential issues, relevant to the test method considerations highlighted in Chapters 3 and 4, will be summarised and discussed.

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2. Nanoparticles and environmental concerns 2.1. What are nanoparticles? Nanoparticles belong to the wider group of nanomaterials, where the prefix ‘nano’ refers to infinitesimal physical dimensions and where particles are defined as a “minute piece of matter with defined physical boundaries” where “physical boundary can also be described as an interface” (ISO, 2008). Many definitions have been proposed for nanoparticles and nanomaterials. Early definitions were based mainly on size, but with an increased understanding of the governing characteristics of nanomaterials, definitions have been refined. Definitions can be divided into two groups: regulatory and scientific. Regulatory definitions are often simplified to encompass the wide range of different nanomaterials in order to provide a common ground for discussion among regulators, industry and public. Some efforts are seen in the scientific literature to define nanoparticles based on their novel size-dependant properties. For example, it has been suggested that for inorganic nanoparticles, only particles of sizes below 30 nm exhibit properties that are different from the corresponding larger (bulk) particle of the same material (Auffan et al., 2009). A common definition of engineered nanoparticles, combining both size and property characteristics, is as particles with dimensions of about 1 to 100 nm, purposefully manufactured to have unique properties (Kreyling et al., 2010; Auffan et al., 2009; National Nanotechnology Initiative, 2007). Hence nanoparticles possess properties that are “qualitatively or quantitatively distinctly different from their other physical forms” (SCENIHR, 2006), such as those of larger-sized particles (bulk particles) made from the same materials and their water-soluble/ionic form. Size-related differences in particle properties may be due to the larger surface area per mass, resulting increased ratio of surface-to-core atoms and increased number of corner and edge atoms. This results in increased reactivity (Feldheim, 2007) or increased ion release (Elzey & Grassian, 2010), which enables their use in novel applications. Furthermore, engineered nanoparticles are classified as a group separate from naturally occurring nanoparticles and anthropogenic incidentally produced nanoparticles (Oberdörster et al., 2005). 2.2. Applications At present, commercial applications of nanomaterials are seen mainly in consumer products and cosmetics (Hansen et al., 2008). In 2009, more than 1000 consumer products based on nanotechnology were registered in the Project on Emerging Nanotechnologies’ inventory; an almost four-fold increase since 2006. According to product descriptions, specific materials could, in some cases, be

http://www.ncbi.nlm.nih.gov/pubmed?term=%22Oberd%C3%B6rster%20G%22%5BAuthor%5D�

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associated with these products, such as silver (Ag) (256 products), carbon including fullerenes (e.g. C60) (82 products), TiO2 (50 products), SiO2 (35 products) ZnO (30 products) and gold (Au) (27 products) (The Project on Emerging Nanotechnologies, 2010). Examples of consumer products containing nanomaterials are stain-repellent and antibacterial textiles, transparent sunscreens, easy-clean surface coatings, sports equipment etc. In an attempt to ensure traceability and transparency, the Belgian EU Council Presidency has advocated a mandatory labelling of consumer products containing nanomaterials (Euractiv, 2010). Yet this proposal has not, so far, resulted in any labelling requirements. Due to recent amendments in the European Cosmetics Directive, however, nanomaterials in cosmetic products have soon to be specified by name followed by the word ‘nano’ in the list of ingredients (Nanotechnology Industries Association, 2009). These new requirements are applicable as of July 2013 (European Commission, 2010). As mentioned above, the primary focus of this thesis is on metal, metal oxide and carbon-based nanoparticles, in particular gold, TiO2 and C60 nanoparticles. The latter are all on the ‘list of representative manufactured nanomaterials’, put forward by the OECD’s Sponsorship Programme for the Testing of Manufactured Nanomaterials and selected for being in (or close to) commercial use and/or produced in high volumes (OECD, 2010a). Metal oxide nanoparticles have been said to represent a “fundamental cornerstone of nanoscience and nanotechnology” (Pinna & Niederberger, 2008) due to their variety of properties and potential applications. However, this also provides evidence of the fact that metal oxides include many and diverse types of nanoparticles with large differences in chemical composition and behaviour; nanoparticles of TiO2, ZnO, CuO and CeO2 comprise some of the more common examples. Besides differences in nanoparticle composition, variations occur in, for example, particle size, shape and crystallinity. On the one hand, this makes them suitable for different applications (such as photonics, energy conversion and storage, catalysis, biomedical applications, healthcare products and self-cleaning surfaces (Sharma, 2009; Pinna & Niederberger, 2008; Nowack & Bucheli, 2007) and, on the other hand, will also lead to differences in their biological effects. In Chapter 3, the ecotoxicity of metal oxides to aquatic organisms is reviewed with particular focus on effects of TiO2 nanoparticles. Manufactured nano-sized TiO2 particles are an example of a nanomaterial already widely used in various applications – many of which are based on their ability to absorb UV-light and their photocatalytic activity – which has been

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found to increase when particle size is decreased (Gao & Zhang, 2001). Generally, TiO2 mainly occurs naturally in its rutile crystalline form (Sharma, 2009). Larger micron-sized TiO2 particles and nanoparticles >10 nm are being produced as both rutile and anatase. Nanoparticles <10 nm are mainly produced as the anatase form, which has been found to be more photocatalytically active compared to the rutile form (Sayes et al., 2006). The combination of anatase crystal structure and small size is therefore likely to alter interactions with living organisms and toxicity, compared to larger-sized particles. Also, ZnO is known to have a high photocatalytic activity under UV illumination, whereas other metal oxides such as CeO2 and CuO display much lower activities (Miyauchi et al., 2002; Yabe & Sato, 2002). Metallic nanoparticles can be composed of a wide range of noble metals such as gold (Au), silver (Ag), platinum (Pt), palladium (Pd), as well as base metals such as nickel (Ni), cobalt (Co) and copper (Cu). Noble metal nanoparticles are well studied in relation to various applications. Some (including Pt and Au) have applications in catalyst materials, whereas others possess antibacterial properties (Ag) or are used in biomedical applications (Au). Particle synthesis can be controlled so that metallic nanoparticles with narrow size distributions can be prepared (Neouze & Schubert, 2008). Methods exist for so-called ‘green synthesis’ which are thought to increase biocompatibility and applicability of metal nanoparticles in biological applications due to the omission of harsh chemicals such as solvents. For example, glucose and starch can be used as the reducing and stabilising agents in the production of stable colloidal suspensions of Ag and Au nanoparticles (Engelbrekt et al., 2009), or metal nanoparticles can be produced by intracellular or extracellular biosynthesis (Thakkar et al., 2009). The surfaces of metallic nanoparticles can easily be functionalised, for example with organosulphur compounds which will spontaneously adsorb to metal surfaces (Neouze & Schubert, 2008). Nanosilver is one of the engineered nanoparticles with most current uses in various applications (Elzey & Grassian, 2010), including for example in textiles, personal care products, food storage containers and laundry additives, due to its antibacterial effects (Navarro et al., 2008a). Au nanoparticles, on the other hand, are considered to be of low-toxicicity and are valued in biomedical applications due to their biocompatibility. They are also resistant to oxidation (Neouze & Schubert, 2008) and have a high density which gives good contrast in various imaging techniques such as transmission electron microscopy (TEM) (Renault et al., 2008). Gold nanoparticles are among the most studied metal nanoparticles in relation to, for example, physiochemical properties and applications (Neouze & Schubert,

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2008), whereas silver is the most studied type of metal nanoparticle when it comes to ecotoxicological effects (Baun et al., 2009). However, the number of studies on the ecotoxicity of Au nanoparticles in the scientific literature is very limited at present. The group of carbon-based nanoparticles comprises both fullerenes and carbon nanotubes (CNTs). Fullerenes are ball-shaped carbon molecules with a cage-like structure that exist with various numbers of carbon atoms. The C60 molecule has been found to exhibit an exceptional physicochemical stability (Kroto et al., 1991) and to possess many interesting characteristics including the ability to encapsulate small molecules such as H2 inside the carbon cage (Komatsu et al., 2005). Due to its unique features, C60 is thought to be the most studied type of fullerene (Andrievsky et al., 2010). C60 nanoparticles are widely available at low cost and high purity (Oberdörster et al., 2006) and have many and diverse potential uses in biological applications and material science (Andrievsky et al., 2010; Bosi et al., 2003). CNTs share the cage-like structure of the fullerenes, but have a tubular shape and a large length/diameter ratio. CNTs can be either single-, double- or multi-walled (termed SWCNT, DWCNT and MWCNT, respectively), depending on the number of carbon shells and have potential uses in applications such as composite materials, hydrogen storage and in electronic devices (Popov et al., 2004). Though present use of nanomaterials might not clearly seem to indicate the start of a “new industrial revolution”, as nanotechnology has been described by many (Kreyling et al., 2010; Bhushan, 2007), future applications are expected to have more widespread benefits in areas as diverse as nanomedicines (e.g. within cancer treatment), groundwater remediation and superconductors (Grieger et al., 2010; Nam et al., 2009; Kang et al., 2006). Nanoparticles are used to give products or formulations new and improved characteristics. As a result of current and future production, they will inevitably be released into the environment where they will pose a potential risk. 2.3. Environmental exposure The route of engineered nanoparticles into the aquatic environment can be via accidental as well as intentional release (e.g. through environmental remediation efforts). Once there, their fate will depend on a number of factors such as presence of natural organic matter (NOM), ionic strength and pH. For example, the presence of NOM has been found to have a stabilising effect and prevent

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particle aggregation1

(Keller et al., 2010; von der Kammer et al., 2010). The potential fate of nanoparticles in the aquatic environment and their interactions with aquatic organisms is illustrated in Figure 1. Predictions of environmental concentrations of engineered nanoparticles are hampered by a lack of understanding regarding their fate, transport and behaviour in natural environments (Gottschalk et al., 2009). This also makes it difficult to predict in which environmental compartments different types of nanoparticles may potentially accumulate and which are the relevant target organisms (Oberdörster et al., 2007).

Figure 1. Possible routes of environmental exposure of aquatic organisms after release of engineered nanoparticles into the aquatic environment (Baun et al., 2008a – Paper VI). Along with prediction of environmental fate, information on potential adverse effects is also required to minimise potential environmental and health hazards; risks resulting from the production, use and disposal of nanoparticles. Risk assessment has to be considered at an early stage in the innovation process. Not only will this help to identify hazardous nanoparticles but it may also facilitate

1 Though it is acknowledged that aggregation and agglomeration are two separate phenomena (Oberdörster et al., 2007), it can be difficult to distinguish the two in practice. The term aggregation is used here to describe the clustering of nanoparticles and therefore covers both agglomeration (coagulation) and actual aggregation (fusing).

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the choice of safer alternatives by modifying particle properties. For example, it has been shown that interactions of gold nanoparticles with biological systems are related to particle geometry (Albanese et al., 2010; Hutter et al., 2010). Hence by identifying nanoparticle properties that determine their toxicity and interactions with biological systems it will be possible to design safer nanoparticles and limit their toxic and ecotoxic effects. 2.4 Currently available guidance for ecotoxicity testing of nanoparticles In order to understand and discuss meaningfully the potential negative environmental effects of engineered nanoparticles – both in a scientific and regulatory context – it is necessary to have the ability to conduct appropriate tests for their hazard potential, hereby generating relevant results. A key player in regulatory health and safety testing of chemicals is the OECD Chemicals Committee. One task within the remit of this committee is the development of standard test guidelines for use primarily in regulatory safety testing (OECD, 2011). To deal with the potential risks of nanomaterials, the OECD Working Party on Manufactured Nanomaterials (WPMN) has been established as a subsidiary body to the OECD Chemicals Committee. The WPMN was established in order to “ensure that the approach to hazard, exposure and risk assessment (of nanomaterials) is of a high, science-based and internationally harmonised standard” (OECD, 2010b). As part of their work, the OECD Sponsorship Programme of Testing a Representative Set of Manufactured Nanomaterials was launched in 2007 with the expected outcome of identifying intrinsic nano-specific properties of nanomaterials. This is seen as a prerequisite for choosing, adapting and/or creating appropriate risk evaluation and management strategies (OECD, 2010c). Ongoing work to develop methods and strategies has resulted in several projects and numerous publications in the Series of Safety of Manufactured Nanomaterials. This includes a Preliminary Review of OECD Test Guidelines for their Applicability to Manufactured Nanomaterials. In this it was suggested that supplementary documents, containing guidance on nano-specific test concerns, might be a better option at present rather than extensive modification of all OECD test guidelines (OECD, 2009). The use of existing test guidelines for regulatory testing will continue until such nano-specific guidance documents or actual test guidelines are available (European Commission, 2008).

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The OECD Sponsorship Programme of Testing a Representative Set of Manufactured Nanomaterials is a project undertaken by the WPNM. It consists of two phases where the first (current) phase is of a more exploratory nature and is “less a data development programme and more a method development programme than the WPMN had originally anticipated” (OECD, 2010a). As part of the programme a ‘Guidance Manual for the Testing of Manufactured Nanomaterials’ (OECD, 2010a) and ‘Preliminary Guidance Notes on Sample Preparation and Dosimetry for the Safety Testing of Manufactured Nanomaterials’ (OECD, 2010b) have been published with the aim of providing guidance to participants in the programme in order to ensure comparable test results among the contributing partners. Both of these documents are “expected to be updated and amended in an iterative manner based upon knowledge accumulation” (OECD, 2010a). The Guidance Manual does provide useful suggestions in relation to toxicity testing of nanoparticles, including what to address in relation to characterisation, environmental fate and behaviour and toxicological endpoints, including methods of dealing with these issues. However, it is also stated in the document that, in the first phase, “there is a preference for exploration of sponsored MN properties (…) rather than developing specific data for risk management purposes” (OECD, 2010a). Hence the document is part of an iterative science-based process of developing appropriate guidance rather than currently being a guidance document for regulatory testing. In addition to nano-specific guidance, a number of guidance documents are currently available providing directions on testing of other chemical compounds besides readily soluble organic chemicals. Though these have not been prepared with nanoparticles in view, they deal with some issues that are relevant in a nano-context. As a supplement to standard guidelines for ecotoxicological tests, specific guidance exists for difficult substances, namely the ‘OECD Series on Testing and Assessment’ No. 23 (OECD, 2000) and ISO 14442:2006 (ISO, 2006). Whereas the OEDC guideline is intended as a general guideline for toxicity testing, the ISO standard deals specifically with algal tests. These guidelines are often mentioned in relation to aquatic testing of the ecotoxicity of nanoparticles, though they have not been prepared with nanoparticles in view and are for that reason not likely to cover nano-related methodological problems satisfactorily. This has also been acknowledged in review of the OECD guidance document on difficult substances (OECD, 2000) by the OECD WPNM, in which it was pointed out that recommendations related to characterisation were particularly inadequate. Nonetheless, these two guidance documents cover

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different aspects of testing ‘difficult’ substances which are relevant to the testing of nanoparticles and may provide inspiration for the development of similar guidance documents relating to the testing of nanomaterials (OECD, 2009). Ongoing regulatory efforts to develop testing guidance run in parallel with scientific efforts to understand the intrinsic properties of nanoparticles and their biological effects. It has been pointed out by the European Commission that a rapid improvement of the scientific knowledge basis is needed to support regulatory work including “Data on toxic and eco-toxic effects as well as test methods to generate such data” (European Commission, 2008). Due to little previous experience in the scientific community in dealing with particles in aquatic test systems, there is still a some way to go before guidance documents, covering all relevant aspects of nanoparticle testing, can be expected to be available. Regulatory guidance will improve concurrently with scientific understanding of intrinsic nanoparticle properties, and it may ultimately be acknowledged that nanoparticles represent a complete paradigm shift whereby actual nano-specific test guidelines are in fact needed.

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3. Ecotoxicity, uptake and behaviour of metal, metal oxide and carbon-based nanoparticles in aquatic test systems Reporting of ecotoxicological tests of nanoparticles in the scientific literature has increased enormously during recent years (Figure 2). Whereas less than 50 open peer-reviewed ecotoxicity studies on environmentally relevant species had been published by 2008 (Baun et al., 2008a – Paper VI), this number has increased dramatically since then. However, the discipline of nano-ecotoxicology can still be characterised as being of a somewhat exploratory nature (Baun et al., 2010). The existing literature does reflect a development towards more targeted studies, but also reveals the need for a very methodical approach to understanding the underlying mechanisms.

Figure 2. Development in literature on ecotoxicological effects for selected nanoparticle types available via a search in ISI Web of Knowledge (searched 02.01.2011). Several reviews of scientific literature are available from governmental institutions, international agencies and in scientific journals. A review of existing literature was recently published by the European Commission in 2010 and covers ecotoxicological literature published before December 2008 (Stone et al., 2010). Other reviews include: Farré et al. (2009), Mouchet et al. (2009), Pérez et al. (2009), Sharma (2009), Baun et al. (2008a – Paper VI), Kahru et al. (2008), Klaine et al. (2008) , Navarro et al. (2008b) and Nowack & Bucheli (2007). Due

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to the rapid expansion in the scientific literature, even the most recent of these review papers and reports does not include a large number of more recent nano-ecotoxicology studies. The following brief review, based on information in Baun et al. (2008a – Paper VI) and Stone et al. (2010), is accordingly brought up to date with recently published findings. Findings in these review documents, and the scientific literature in general, illustrate that, while general conclusions cannot be drawn relative to nanoparticle ecotoxicity, due to the great diversity in material types and particle properties, some types of nanoparticles have been found to cause acute and sub-lethal ecotoxic effects at concentrations in the μg/L range, whereas others have low or no toxicity at mg/L concentrations. In many studies it has been possible to establish dose-response relationships but large variations in effect concentrations have been found even in seemingly comparable tests using nanoparticles of the same material and the same test species. The purpose of this review is to investigate the causes of the observed effects as well as possible reasons for the large discrepancies. It is focused on specific issues of interest – either related to particular types of nanoparticle or of general concern in the field of nanoecotoxicology as a whole. Reliability and interpretation of published findings will be discussed in relation to possible toxic mechanisms and impeding factors. 3.1. Direct and indirect effects of metal oxide nanoparticles in algal growth inhibition tests Ecotoxicological effects of a number of metal oxide nanoparticles have been investigated for several freshwater algal species, with Pseudokirchneriella subcapitata as the most common test organism. Generally, metal oxide particles have been found to form larger aggregates >100 nm in both deionised water and fresh water (e.g. Bai et al., 2010; Brayner et al., 2010; Hartmann et al., 2010 – Paper I; Keller et al., 2010). Several studies have found nano-sized metal oxide particles to be more toxic to algae compared to larger-sized bulk particles (e.g. Hartmann et al., 2010 – Paper I; Rogers et al., 2010; Wei et al., 2010; van Hoecke et al., 2008) when dose is expressed as mass and EC50 values are generally in the mg/L range (e.g. Hartmann et al., 2011b – paper IV; Menard et al., 2011; Wei et al., 2010; Hartmann et al., 2010 – Paper I). Efforts have been focussed on trying to establish links between effects, particle characteristics and behaviour in the test systems. More specifically, the effect of different sizes of the same material, development in aggregation behaviour over time and the physical interactions with algal cells have been investigated. However, it is

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unlikely that there will be a simple descriptor for the toxic potential of metal oxide nanoparticles to algae due to the large variations in particle properties and behaviour. Ion-releasing nanoparticles, such as ZnO, represents a special case for which toxicity has been attributed (at least partly) to the release of Zn2+ ions (Poynton et al. 2011; Bai et al., 2010; Aruoja et al., 2009; Franklin et al., 2007), whereas dissolution of metal ions from, for example, TiO2 and CeO2 nanoparticles has been found to be negligible (Johnston et al., 2010; Rogers et al., 2010). Adsorption or clustering of metal oxide nanoparticles to/around algal cells has been observed in several studies and demonstrates an affinity of algal cell surfaces for nanoparticle attachment (Hartmann et al., 2011b – Paper IV; Brayner et al., 2010; Hartmann et al., 2010 – Paper I; Aruoja et al., 2009; van Hoecke et al., 2009) (Figure 3). Nano-sized particles has been observed to cover algal cell surfaces to a greater extent than corresponding larger-sized particles for which a greater number of particle-free algal cells were seen (Aruoja et al., 2009). Actual internalisation of metal oxide nanoparticles into algal cells has at present, however, only been observed for ZnO nanoparticles (Brayner et al., 2010). The close interaction between algal cells and nanoparticles potentially makes several effect mechanisms plausible. These can be divided into indirect effect mechanisms, caused by changes to the cell environment (changing pH, redox, nutrient or light conditions), or direct effects caused by direct physical or chemical interactions between nanoparticle and cell (Rogers et al., 2010). Even if changes in indirect parameters, such as light and nutrient availability, are monitored, it has been pointed out that changes in the proximate cell environment might differ from overall conditions (Hartmann et al., 2010 – Paper I; van Hoecke et al., 2009). Though several studies have been designed to elucidate the possible shading effect of nanoparticles in algal growth tests, most of these have applied a physical separation of algal cells and nanoparticle suspensions (Hartmann et al., 2010 – Paper I; Aruoja et al., 2009; van Hoecke et al., 2009; Hund-Rinke & Simon, 2006.). By using this method, shading has been rejected as a cause of the observed reductions in growth. However, such test designs cannot take into account possible shading taking place at a cellular level, which can be described as an indirect effect caused by the encapsulation of the cells by particles, changing the physical growth conditions (light intensity) (Hartmann et al., 2010 – Paper I). One study has shown that after 96 hours exposure of Scenedesmus

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obliquus to SiO2 nanoparticles (10–20 nm, 25-200 mg/L) cell contents of chlorophyll decreased in a concentration-dependent manner, whereas the content of carotenoid was unaffected. This was hypothesized to indicate shading on a cellular level (Wei et al., 2010). Content and composition of carotenoid is known to be affected as a result of photoacclimation (Dubinsky & Stambler, 2009). Hence, detailed investigations of such changes may assist to elucidate the effect mechanisms of nanoparticles towards algae.

Figure 3. Encapsulation of algal cells (Pseudokirchneriella subcapitata) exposed to TiO2 nanoparticles. A: Scanning electron microscopy (SEM) image of an algal cell after exposure to 50 mg/L TiO2 in ISO algal test media for 48 h. B: Corresponding SEM-EDX dot map shows the distribution of Ti. It can be seen that the TiO2 nanoparticles cover the surface of the algae (Modified from Hartmann et al., 2010 – Paper I). C: Transmission electron microscopy (TEM) images showing the formation of algae-particle heteroaggregates (scale bar: 2 μm). (Modified from Hartmann et al., 2011b – Paper IV) Particle adhesion may also lead to direct physical effects such as disruption of cell membrane as a result of their surface structure or due to photochemical reactions taking place on or near the cell surface. It has been suggested that nanoparticles may cause physical disruptions of cell walls and membranes as a result of their topography and surface properties (Hartmann et al., 2010 – Paper I. Stevens & George, 2005; Andersson et al., 2003 among others). The nano-scale roughness is thought to increase contact between cells and nanoparticles and hot

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spots of radical species could be present at the tip of particle edges (Yeung et al., 2009). Hence this might to some extent explain observations of increased toxicity of nanoparticles compared to their bulk counterparts despite aggregation. CeO2 nanoparticles (10-20 nm) have been found to increase cell membrane permeability of P. subcapitata and the effect was more pronounced for nano-sized particles than for the corresponding bulk particles (<5μm). The effect was believed to require physical interaction between cells and particles and to result from photochemical reactions (Rogers et al., 2010). Comparable results have been obtained by Rodea-Palomares et al. (2011) who observed cell wall disruption and leaking of cytoplasm after exposure to CeO2 nanoparticles with primary particle sizes of 10–60 nm. Again it was hypothesised that the effect mechanism was related to a direct contact between nanoparticles and cells. Similarly, ZnO nanoparticles have been found to damage the cell wall structure, cause leaking of intracellular content and internalise by endocytosis in Euglena gracilis (Brayner et al., 2010). Without particle adhesion, nanoparticles in the test systems could also potentially lead to other indirect effects by reducing availability of nutrients such as phosphate (Rogers et al., 2010; van Hoecke et al., 2009). Other effect mechanisms include the production of reactive oxygen species (ROS), which may both be extracellular and intracellular. As a result of the photocatalytic activity of TiO2, ROS (including hydroxyl radicals) are known to be generated. These are very potent oxidants with non-selective reactivity. Though they are short-lived, nearby cell surfaces are potential targets resulting in lipid peroxidation, membrane damage, increased membrane permeability and hereby decrease cell viability of E. coli (Maness et al., 1999) as well as planktonic freshwater organisms (Battin et al., 2009). Production of intracellular ROS is another potential effect mechanism of TiO2 nanoparticles, which has possibly been observed by Battin et al. (2009) in natural microbial communities. However, intracellular uptake of TiO2 nanoparticles, or of extracellular generated ROS entering the cell, could also explain the observed effects, which were detected by means of a cell permeable ROS indicator (2′,7′-dichlorofluorescein diacetate) (Battin et al., 2009). When exposing test organisms to a non-dissolved test substance, separation of physical effects from actual toxic effects can be difficult (OECD, 2000). When testing the effects of engineered nanoparticles this is precisely the case – and presents a major challenge. As can be seen from the research discussed here, for some types of metal oxide nanoparticles the underlying mechanisms causing the

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observed effects have been elucidated through studies that have included several additional endpoints compared to standard ecotoxicity tests. This illustrates the need to move beyond standard test systems in order to gain a deeper insight, which may prove useful in guiding the way forward for ecotoxicological testing of nanoparticles. 3.2. Intracellular uptake in algae It is well known for mammalian cells that particle geometry and surface alter cellular uptake and that spherical nanoparticles, especially those in the range of 40-60 nm, can be readily internalised and accumulated (Albanese et al., 2010). For example, it has been shown that interactions of gold nanoparticles with biological systems are related to particle geometry (Hutter et al., 2010). Also surface charge (controlled by functionalisation) has been found to determine cellular uptake as positively charged gold nanoparticles are taken up by different malignant and nonmalignant cell types to a far greater extent than neutral and negatively charged particles of the same size (~ 2 nm core size) (Arvizo et al., 2010). As opposed to human and animal cells, algal cells have a cellulose cell wall, which acts as an additional barrier to particle uptake. The cell wall is semi-permeable with pores in the size range of 5-20 nm through which nanoparticles must pass in order to reach the plasma membrane. However, nanoparticles may in themselves be able to induce pore formation which may, in turn, allow for uptake of larger-sized particles (Navarro et al., 2008b). Cell wall disturbances in green algae by gold nanoparticles have been proposed as a potential effect by Renault et al. (2008). Actual internalisation into algal cells has been observed in species known to be capable of endocytosis. For example, ZnO nanoparticles have been found in the cytoplasm of the heterotrophic alga Euglena gracilis. Uptake was in one case found to be via endocytosis and the uptake mechanism varied depending on the type of protective agent used in synthesis to stabilise the nanoparticles. (Brayner et al., 2010). Also Ag nanoparticles have been found to internalise in Ochromonas danica. This species of alga is able to take up bacteria and yeast through phagocytosis and utilise these as an additional carbon source. Results showed that addition of gluthatione to mediate the action of Ag+ did not eliminate the latter completely, indicating an effect of the internalised nanoparticles (Miao et al., 2010). Compared to photoautotrophic algae, such as P. subcapitata, these heterotrophic and mixotrophic algae might therefore be more susceptible to nanoparticle exposure as nanoparticle uptake is more likely to take place.

http://en.wikipedia.org/wiki/Autotrophic�

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3.3. Particle versus metal ion effects The toxicity of the soluble forms of metals provides an indication of the possible toxic effects of the same metals produced as a nanoparticulate form. It has been found that metallic and metal oxide nanoparticles composed of elements that are in themselves toxic (such as Cu, Zn and Ag) show higher ecotoxicity than other types of metal nanoparticles composed of less toxic elements (Kahru & Dubourguier, 2010; Griffit et al., 2008). Furthermore, it has been concluded by several studies that toxicity of some nanoparticles (such as Ag) can partly or largely be explained by the release of free ions (e.g. Ag+) (Miao et al., 2010; Miao et al., 2009; Navarro et al., 2008a). The degree to which ions are released from nanoparticles varies depending on composition. It has been found that release of Ti and Ce from TiO2 and CeO2 nanoparticles is very low (<10 μg/L) (Johnston et al., 2010), whereas the release of Zn from ZnO nanoparticles can be in the mg/L range (Franklin et al., 2007). The increased toxicity of CuO nanoparticles to algae compared to bulk counterparts has been related to increased ion release as a 141-fold increase in available Cu was seen for the nano-sized particles (30 nm) compared to larger particles (size not stated) (Aruoja et al., 2009). Ag is known for its antimicrobial properties and has been found to induce ecotoxicological effects. In one study it was found that the toxicity of Ag nanoparticles to the freshwater alga Chlamydomonas reinhardtii was caused by the release of Ag+. Furthermore, results showed that algae could mediate the release of Ag+ from the nanoparticles (Navarro et al., 2008a). However, other studies have indicated that the explanation of Ag nanoparticle effects may not be so straightforward (Miao et al., 2010; Griffitt et al., 2008). In one study, only 0.07% of the total Ag nanoparticle mass was found to dissolve and its toxicity towards zebra fish (Danio rerio) and Daphnia pulex was found to be due instead to a particle-related effect. The same study also found that D. pulex and Ceriodaphnia dubia were more sensitive to nano-metal exposure than algae and D. rerio. It is hypothesised that their filter-feeding behaviour might be the reason for this (Griffitt et al., 2008). It is plausible that the large uptake of nanoparticles by filtration, which will be described in Section 3.5., may lead to an intra-organism source and release of metal ions causing toxicity. This will not be detected by measuring the ion concentration in the surrounding media. The toxicity of ZnO nanoparticles to the freshwater green alga P. subcapitata was studied by Franklin et al. (2007), who suggested that the negative effects could be attributed to dissolved Zn(II) originating from the ZnO nanoparticles.

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Thus, toxicity was not directly related to the nanoparticulate form of ZnO, but rather to the presence of ZnO in the tests. A similar conclusion was reached by Aruoja et al. (2009). However, by looking into changes in Daphnia magna gene expression, Poynton et al. (2011) found that exposure to ZnO nanoparticles resulted in gene expression profiles which differed from those of daphnids exposed to ZnSO4, suggesting that toxicity is related both to Zn2+ ion release and an actual particle effect. This has been supported by Bai et al. (2010), who found that the release of Zn2+ from ZnO nanoparticles (~40 nm) could not completely explain the decreased hatching rates of D. rerio embryos. Effects on body length and malformations were also more severe for ZnO nanoparticles compared to corresponding concentrations of Zn2+. The contribution to toxicity from free ions and particles, respectively, is clearly something that should be investigated further. DNA microarrays seem to have a potential role in this work as they allow for a differentiation between particle and ion effects. Also luminescent bacterial tests have been applied to investigate effect contributions from both ion release and ROS generation to the toxic effects of Ag, ZnO, TiO2 and CuO nanoparticles (Ivask et al., 2010). A way to investigate actual particle-related effects is to test inert particle types. If the core material is in itself inert, particle-related effects can be investigated as a function of, for example, size and functionalisation. As a material, Au is generally seen as being inert and as offering possibilities for narrow size distribution and functionalisation. For these reasons, Au nanoparticles could be a good candidate for a model nanoparticle, and it is also included on the OECD WPNM list of representative nanomaterials (OECD, 2010d). However, apart from that reporting the use of algae and other microorganisms in so-called green synthesis of gold nanoparticles, not much literature exists regarding biological interactions between gold nanoparticles and algae. In one study, the effects of amine-coated 10 nm gold nanoparticles were investigated using the freshwater alga (Scenedesmus subspicatus) and the benthic bivalve (Corbicula fluminea) as test organisms. Results showed that amine-coated Au nanoparticles (10 nm) reduced algal cell density. Furthermore, the cell size of exposed cells was found to be decreased. Au nanoparticles were strongly adsorbed to the algal cell wall and thought to cause intracellular and cell wall disturbances. In the same study, C. fluminea was exposed to gold-contaminated algae and Au nanoparticles were observed to penetrate branchial and digestive epethelia (Renault et al., 2008). EC50 values for P. subcapitata exposed to starch-coated 25 nm Au nanoparticles for 72 hours were found to be >48 mg/L based on visual cell counts. However, after 48 hours exposure pigment content did not increase further in algae cultures

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exposed to >6 mg/L. This indicates an effect on pigment synthesis though the underlying mechanisms remain to be clarified (Hartmann et al., 2011b – Paper IV). 3.4. Effects resulting from impurities and solvents In mammalian toxicology studies, length and degree of entanglement have been found to be determining factors for the effects of MTCNTs, where long non-tangled MWCNTs induced asbestos-like mesothelial injuries (inflammation and granuloma formation) (Poland et al., 2008). However, such correlations have not, as yet, been seen in ecotoxicological tests. Test results of CNT toxicity can be hampered by the presence of metal impurities, since metals (e.g. Co, Ni and S) can be found in trace concentrations (Cheng et al., 2007) in the order of 3–10% (Cañas et al., 2008). Variations in uptake of SWCNT in the estuarine copepod Amphiascus tenuiremis have been thought to be size-dependent and due to differences in nanoparticle aggregation behaviour (Templeton et al., 2006). Underivatised MWCNTs were found to have LC50 values of 17mg/L [14–20]95% and 21mg/L [18–24]95% for the crustacean C. dubai when prepared through either stirring or sonication (Kennedy et al., 2009). Roberts et al. (2007) have demonstrated a dose-dependent acute toxicity of lysophosphatidylcholine-coated SWNTs to D. magna. After 96 hours of exposure, a mortality of 20% and 100% was seen for organisms exposed to 10 and 20 mg/L, respectively. The organisms were able to use the coating as a food source, hereby biologically modifying the CNTs and increasing precipitation. NOM-stabilised MWCNTs have also shown toxicity to D. magna and reduced the growth and reproduction of C. dubai. For D. magna, effects were attributed to ingestion of MWCNTs, leading to clogging of the digestive tract. It was further observed that, despite disaggregation of MWCNT in the digestive tract, no transport of particles across the gut lumen took place (Edgington et al., 2010). In general, investigations of the ecotoxicological effects of CNTs are complex due to the variation in CNT characteristics, impurities and bundling behaviour and precise determining characteristics for the aquatic effects of CNTs have not been identified at present. These are clearly all issues which need to be addressed when analysing results obtained in ecotoxicity tests of CNTs. The ecotoxic effects of C60 fullerenes has been heavily debated in the scientific literature since the publication of the first ecotoxicological study in 2004 in which a significant increase in lipid peroxidation in the brain of juvenile largemouth bass (Micropterus salmoides) was demonstrated after exposure to

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aqueous suspensions of C60 (termed nC60) (Oberdörster, 2004). Studies on the aquatic ecotoxicity of nC60 have been carried out on a variety of different organisms, including crustaceans, algae, bacteria and fish. In many of these studies, the main topic of debate is related to the causes behind the observed effects. The use of tetrahydrofuran (THF) in the preparation of the nC60 has led to discussions about whether the observed effects were caused by the C60 itself or were due to the presence of THF. nC60 generally causes less adverse effects when prepared by stirring and without the use of a solvent, compared to when nC60 is prepared using THF. Low toxicity to D. magna of nC60 has been found with lethal concentrations generally first occurring in the mg/L range (Lovern & Klaper, 2006). Long-term exposure of D. magna to 2.5 mg/L water-stirred nC60 significantly reduced the number of offspring, delayed moulting and increased cumulative mortality (Oberdörster et al., 2006). When THF has been used in the preparation of nC60, acute and chronic effects have been observed at lower concentrations. Examples of this include Zhu et al. (2006) and Lover & Klaper (2006), who both found LC50 values for D. magna <1 mg/L, which corresponds to a significant increase in the toxicity of nC60THF compared to nC60aqu. Characterisation of nC60 suspensions has revealed that the use of THF in their preparation results in smaller aggregates compared to water-stirred nC60 (Spohn et al., 2009). Based on further studies, it has been concluded that the effects of nC60THF on D. magna (Spohn et al., 2009) and D. rerio (Henry et al., 2007) were caused by by-products resulting from THF oxidation in the TFH-nC60 preparation procedure. However, as it has also been pointed out by Tao et al. (2009), the use of THF in the preparation procedure also results in smaller-sized aggregates (10-20 nm) compared to, for example, sonication (20-100 nm) (Lovern & Klaper, 2006) or prolonged stirring (30-600 nm) (Tervonen et al., 2010; Baun et al., 2008b – Paper VII), which may also influence differences in toxicity. These findings emphasise not only the need for thorough particle characterisation but also the care required in sample preparation so that preparation artefacts are not mistaken for toxic effects of nanoparticles. 3.5. Uptake of nanoparticles into the gut of filter and sediment feeders Several studies have described the uptake, mainly of TiO2, Au and carbon nanoparticles, in two species: D. magna and Lumbriculus variegatus. Both these organisms have an apparently non-selective uptake of particles from the water phase and/or sediment. The natural habitat of L. variegatus is mainly shallow waters, where it feeds on decaying vegetation and microorganisms. Studies have

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shown that L. variegatus ingests particles of 40-60 µm in size, whereas larger particles (100-300 µm) are too large relative to the size of its mouth (Miño et al., 2006; Lawrence et al., 2000). No information on the lower size limit for uptake has been found in the literature. By filtration of water, D. magna catch particles (including mainly algae) in the size range 0.4–40 μm (Gophen & Geller, 1984; Geller & Müller, 1981). As a result of this non-selective uptake it is likely that nanoparticle aggregates will also be taken up by these organisms. Besides uptake of pure nanoparticle aggregates, ingestion of particles attached to algae is also a probable uptake route. Uptake by other routes is also possible, but would require transport across cell membranes directly from the water phase, for example as a result of particle adhesion to the organism surface and cell membrane disruption. Studies have been carried out to investigate the uptake and translocation of Au, TiO2, CNTs and C60 nanoparticles in D. magna (Hartmann et al., 2011c – Paper V; Tervonen et al., 2010; Zhu et al., 2010; Petersen et al., 2009; Baun et al., 2008a – Paper VI; Baun et al., 2008b – Paper VII; Lovern et al., 2008) and L. variegatus (Hartmann et al., 2011c – Paper V; Petersen et al., 2008). Hartmann et al. (2011c – Paper V) found that TiO2 nanoparticles were taken up by L. variegatus, indicating that nanoparticles aggregates are at least partly within a size range that makes them available for ingestion (Figure 8). Even though single nanoparticles are also evidently smaller than the lower size limit for D. magna uptake through filter feeding, nanoparticle aggregates are seen in the gut of D. magna after exposure to TiO2 and C60 (Figure 8). This corresponds to general results showing that aggregates in freshwater suspension are micron-sized rather than nano-sized. This is also in agreement with studies by Tervonen et al. (2010) observing nC60 aggregates in sizes of 1100±500 nm inside the gut of D. magna. However, uptake of smaller nanoparticle aggregates has also been observed by Lovern et al. (2008), demonstrating that single Au nanoparticles (17-23 nm) were present in the gut of D. magna after six hours exposure. As a means of facilitating digestion, daphnids are known to ‘drink’ water (Gills et al., 2005) and uptake of such small particle sizes can be a result of direct deposition from the surrounding media (Rosenkrantz et al., 2009). Studies have shown that nC60 and MWCNTs are, to a large extent, taken up into the digestive tract of D. magna after 24-48 hours exposure (Tervonen et al., 2010; Zhu et al., 2010; Petersen et al., 2009; Baun et al., 2008a – Paper VI; Baun et al., 2008b – Paper VII) (Figure 4).

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Figure 4. Uptake and adhesion of nanoparticles in crustaceans. A1 & B1: C60 and TiO2 is seen in the digestive tract of Daphnia magna after 48 hour exposure to 3 mg/L nC60 and 40.5 mg/L TiO2 prepared by prolonged stirring in indirect sunlight and sonication respectively. A2, B2 & C2.: Adhesion of nanoparticles to the exoskeleton of D. magna and Acartia tonsa (Modified from Baun et al. Scale bars correspond to 200 μm. (2008a – Paper VI)). Adverse effects were suggested as being related to disturbance of the function of the digestive tract by nanoparticle packing between the microvilli of the gut. Additionally, the extra energy-use required to excrete the particles trapped between the villi could cause sub-chronic and chronic effects (Tervonen et al., 2010; Petersen et al., 2009). In one case, depuration was found to be a relatively rapid process, but with residual nC60 left in the digestive tract of D. magna (Baun et al., 2008b – Paper VII), whereas slower excretion was observed by Tervonen et al. (2010). In another study, the organisms were found unable immediately to excrete MTCNTs (Petersen et al., 2009). For ingested sediments, clearance from the gut of D. magna has been found to be limited, even after 48 hours (Gillis et al., 2005), so this is not likely to be a nano-specific behaviour. Translocation of 20 nm polystyrene nanoparticles from the digestive tract into lipid droplets of the daphnids has, however, been observed by Rosenkrantz et al. (2009). Au nanoparticles might also have been taken up into the tissue of a microvillus in D. magna (Lovern et al., 2008).

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The potential for bioaccumulation of an organic chemical can be predicted by its water-octanol partitioning coefficient, Kow. The underlying assumption is that octanol is an appropriate representative for lipids in the organisms, also implying that tissue with high lipid content will have increased concentration levels. For uptake of metals, the mechanisms are more complex and may be described by the ‘Free Activity Ion Model’ (FIAM) or the ‘Biotic Ligand Model’ (BLM). Both of these models are based on the assumption that the concentration of available reactive species of a metal is directly linked to metal bioaccumulation and hence to biological effects (Worms et al., 2006). In the process of bioaccumulation there are normally considered to be three general uptake mechanisms (through respiration, ‘dermal’ diffusion and ingestion) and six clearance mechanisms (respiration, ‘dermal’ diffusion, excretion, metabolic conversion, reproduction and dilution by growth) (Mackay & Fraser, 2000). For nanoparticles, on the other hand, uptake in organisms with non-selective particle intake from the surrounding media, ingestion and excretion are likely to be the predominant mechanisms. Furthermore, bioaccumulation of ‘conventional’ chemicals is related to passage of biological membranes, mainly through passive diffusion or active uptake, such as transport through ion channels or carrier mediated transport (Sijm et al., 2007). For nanoparticles, other tissue uptake mechanisms are relevant such as phagocytosis. Additionally, in the case of ion-releasing nanoparticles, active transport of metal ions is of course also important. For Ag nanoparticles, Zhao & Wang (2010) found that uptake rates were biphasic, indicating that the biokinetics were different for lower and higher exposure concentrations. It was hypothesised that, at higher concentration, the observed higher uptake rates were due to particle ingestion, whereas uptake at lower concentrations were dominated by first-order uptake kinetics (Zhao & Wang, 2010). A drawback of this study is the lack of investigations of aggregate sizes in the media. If a larger degree of aggregation takes place at higher concentrations this might explain the differences in uptake through filtration. Nanoparticle uptake into the gut will be especially significant for those types that tend to form larger-sized aggregates in appropriate size ranges. This further entails that uptake of nanoparticles does not necessarily make them bioavailable in the sense that they will reach target organs and cause biological effects. Hence investigations of additional uptake routes, their ability to cross digestive epithelia and translocation to other parts of the organisms (such as described by Rosenkrantz et al. (2009)) might therefore be more relevant than measurements of total body burden. Also, it might be appropriate to differentiate between body burden (defined here as the total organism’s particle load including gut contents) and actual uptake into the tissue.

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In a study by Zhu et al. (2010) bioconcentration factor (BCF) values > 5x104 (based on dry weight) were determined based on steady-state concentrations for TiO2 nanoparticles in D. magna. Similarly, an uptake of 4 to 5 g C60/kg wet weight has been found by Tervonen et al. (2010) with no statistically significant difference between 0.5 and 2 mg/L exposure concentrations. This corresponds to a BCF value of 2000 (2 mg/L C60 exposure) and 7600 (0.5 mg/L C60 exposure). When comparing this to Figure 4, where the presence of nanoparticles in the digestive tract of D. magna is clearly seen, such very high BCF value estimates are not surprising. Though BCF values based on total concentrations (and mainly related to the gut content) will give a conservative result, it may very well be seen as being overly cautious to define, for example, TiO2 nanoparticles as very bioaccumulative based on such results. Whether gut content should be taken into account when determining bioaccumulation is under debate. It has for example been argued by Petersen et al. (2009) that the term BCF should not be used when the nanoparticles are not taken up into the tissue. As has been described by van Geest et al. (2010), contaminants associated with gut content results in an overestimation of tissue concentrations. This is a common procedure for sediment-dwelling organisms. Purging the gut contents prior to preparation for chemical analysis is a way of minimising this error but the drawback is obviously that part of the contaminant associated with this tissue is removed in this process. Also, it may underestimate the potential transfer from one trophic level to another (van Geest et al., 2010). In D. magna, the addition of food (algae) has been found to increase gut depuration of TiO2 and Ag nanoparticles resulting in a fast initial clearance period of a few hours followed by a slower depuration phase (Zhao & Wang, 2010; Zhu et al., 2010). This is similar to what has been found for metal-contaminated sediments, where addition of algae resulted in rapid excretion. Gut fullness decreased from ~60 to ~20 % within eight hours clearance. After >8 hours only ~2% of the gut content was estimated to be sediment; the remaining part constituted algae. In comparison, no decrease in gut content was observed in the absence of algae after 48 hours (Gillis et al., 2005). Adding algae may therefore be a way to accelerate excretion and determine residual concentrations of nanoparticles, reflecting their actual uptake into the tissue. In summary, mechanisms of bioaccumulation of nanoparticles in D. magna and L. variegatus are different compared to ‘regular’ organic chemicals and metals. Total uptake might not be directly related to specific nanoparticle properties (such as chemical composition), or to concentration per se, but could rather be controlled by the fraction of particles occurring in sizes appropriate for non-

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selective uptake and/or with properties resulting in deposition directly from the water phase (which may not increase linearly with increasing concentration). Exposure volume has been identified as a determining factor because increasing volumes of CNT suspensions caused significantly increased uptake of CNTs in D. magna (Petersen et al., 2009). Furthermore, it has been suggested by Tervonen et al. (2010) that gut volume may be the limiting factor for nanoparticle accumulation in D. magna. By comparing the limited number of available studies on nanoparticle uptake in D. magna (assuming 8% dry weight to wet weight ratio for conversions), it seems that nanoparticle uptake increases with exposure concentration and exposure volume, reaching an upper limit around 50-60 mg/g dry weight (i.e. 5-6% w/w) somewhat independent of particle type (Tervonen et al., 2010; Zhao & Wang, 2010; Zhu et al., 2010; Petersen et al., 2009). It should, however, be pointed out explicitly that this observation is based solely on four studies testing four very different particle types (TiO2, C60, Ag and MWCNT) in concentrations up to 2 mg/L. In general, studies on the uptake of nanoparticles in filter feeding and sediment dwelling freshwater organisms show that nanoparticles are taken up mainly into the gut. High quantities of nanoparticles in the gut of organisms may disturb the function of their digestive system. These high levels of non-selective uptake may lead to a much higher body concentration than the surrounding water (Baun et al., 2008b – Paper VII) which may, in turn, also have an effect on higher trophic levels. Transfer of TiO2 nanoparticles from D. magna to C. rerio has been seen but actual biomagnification was not observed (Zhu et al., 2010). Nonetheless, residual concentrations might be more relevant in relation to explaining effects within one tropic level. For ion-releasing nanoparticles, the large uptake may have the consequence that organisms are exposed to increased concentrations of free ions compared to what is predicted from the water-phase concentration. Signs of translocation also herald a warning. If translocation of even a small fraction of the total nanoparticle uptake is then possible into other parts of the organisms, this may eventually lead to toxic effects (Rosenkranz et al., 2009; Baun et al., 2008b – Paper VII). Hence information on translocated mass relative to total body burden (a ‘Translocation Factor’ value) might be a relevant parameter as it will give information on the nanoparticle’s ability to cross the gut lining of various organisms. This is of specific relevance for nanoparticles due to their ‘accidental’ uptake into the gut of freshwater organisms. Methods for the detection of bioaccumulation and translocation, which have been successfully applied, include radio labelling (Petersen et al., 2009; Petersen et al., 2008), neutron activation of metal and metal oxide nanoparticles (Oughton et al., 2008),

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micro-X-ray fluorescence (μXRF) (Hartmann et al., 2011c – Paper V) and TEM (Lovern et al., 2008). In addition to uptake into organisms, adhesion of nanoparticle aggregates to the exoskeleton of crustaceans is also described in the literature (Baun et al., 2008a – Paper VI; Lovern et al., 2007; Roberts et al., 2007) as illustrated for D. magna and Acartia tonsa in Figure 4 (A2-C2). Particle adhesion could physically impair the test organisms and cause behavioural changes. Such physical effects have been discussed by Lovern et al. (2007) and could be an explanation for reduced mobility (Baun et al., 2008a – Paper VI). 3.6. Biomodification of nanoparticle characteristics and behaviour Though biotic and abiotic transformations of nanoparticles may take place, actual biotic degradation is not possible for many of them due to their often inorganic chemical composition. Also carbon-based nanoparticles, such as aged aquatic suspensions of water-stirred nC60, can be classified as not readily biodegradable according to the OECD test procedure (Hartmann et al., 2011a – Paper III), although degradation has been observed under different conditions (oxidised by white rot basidiomycete fungi to CO2 (Schreiner et al., 2009)). This makes it more relevant to investigate other forms of biological modification to particle characteristics and behaviour, degradability of coating materials and abiotic transformations. In algal tests, or when internalised by different invertebrates, biological modifications of nanoparticles have been observed. Some studies have looked into biomodification of nanoparticles as a result of ingestion (Baun et al., 2008b – Paper VII; Griffitt et al., 2008; Roberts et al., 2007).Changes in particle characteristics and behaviour have, in some cases, resulted from the removal of coating layers. Roberts et al. (2007) found that upon ingesting lipid-coated SWCNTs, daphnids were able to remove the coating material, utilise it as a food source and excrete non-coated insoluble SWCNTs. Conversely, Baun et al. (2008b – Paper VII) found a tendency towards decreased aggregate size of nC60 following ingestion and excretion by D. magna (Figure 5). As will be described in section 5.1, it is well-known that changes in the pH and ionic strength of aquatic media will influence the behaviour of nanoparticles. In the same way, ingestion and exposure to the intra-organism environment may result in changes in pH and ionic strength which, in turn, can lead to changes in nanoparticle aggregation (Hang et al., 2009) or ion release (Lui & Hurt, 2010), hereby also possibly affecting particle toxicity and uptake into the organism. The

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internal pH of organisms can vary considerably from that of the surrounding environment. For example, the pH of the digestive tract in D. magna was found to vary from 6.8 at the anterior to 7.2 at the caudal end (Hasler, 1935). This represents the lower end of recommended pH values for test media (6-9) in OECD test guidelines for acute and chronic toxicity (OECD, 2008; OECD, 2004).

Figure 5. Distribution of Phenanthrene between different phases and particle size fractions during exposure to D. magna in the presence of nC60 aggregates. Due to large sorption (>85%) of phenanthrene to nC60, the distribution of phenanthrene also gives an indication of nC60 size fractions (Modified from Baun et al., 2008b. – Paper VII). In algal growth inhibition tests, changed particle size distributions for starch-coated Au nanoparticles were observed during the 72 hour test period (Figure 6). Initial size measurements (using Nanoparticle Tracking Analysis) showed that the particles were relatively mono-dispersed in the solution with hydrodynamic diameters around 50 ± 20 nm. After 24 hours, investigations of the particle size distribution indicated a small decrease in particle size, possibly resulting from degradation of the starch coating layer. However at 48 hours, a very limited fraction of particles could be detected below 2 μm. This indicated either formation of larger-sized aggregates or adsorption onto a growing biomass consisting either of algae or other microorganisms thriving in the advantageous light and temperature conditions, combined with relatively high starch concentrations (Hartmann et al., 2011b – Paper IV). It has been found that exudates of P. subcapitata influence aggregation of colloidal particles as well as reducing metal toxicity (Koukal et al., 2007). Decreased toxicity of Ag nanoparticles resulting from concentration-dependent algal production of exudates in response to nanoparticle exposure has also been reported by Miao et

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al. (2009). Organic matter excreted from algae and other test organisms is therefore also likely to influence nanoparticle behaviour as are also the effects of metal ions released from some nanoparticles. The toxicity of TiO2 nanoparticles to C. dubai has been found to be decreased dramatically by the addition of NOM; EC50 values increased from 3-15.9 mg/L without to >100 mg/L with NOM (Hall et al., 2009). In a study by Griffitt et al. (2008) it was found that Ag and Cu nanoparticles were lost from the water column during exposure (>90 and 50%, respectively, after 48 hours) and that the degree of loss differed between exposures to D. pulex and D. rerio. It was speculated that this was caused by differences in the organic matter excreted by the two organisms. This then leads to a time-dependent and dynamic interaction between organisms and nanoparticles (Nel et al., 2009). However, the influence of organic matter on nanoparticle behaviour is an area of extensive research and the presence of natural organic matter has also been found to have a stabilising effect on nanoparticle suspensions, as will be discussed in Section 5.1.

Figure 6. Development in Au nanoparticle size distributions over a 72 hour incubation period in an algal growth inhibition test. The inserted graph shows size distribution in the range from 0 to 2 μm. The samples (1.9 mg/L) were incubated on a shaking table at 20 ± 2oC and continuously illuminated at 86-109 μE/m2/s. Samples only containing Au nanoparticles were incubated along with algal cultures exposed to the nanoparticles. (Hartmann et al., 2011b – Paper IV) Although test systems that are not only controlled by physiochemical conditions but also actively affected by the test organisms are not an unknown phenomenon relative to ‘conventional’ water soluble chemicals (Newman & Unger, 2003), this clearly presents a special challenge when testing nanoparticles. If their

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coating layer consists of organic materials such as lipids and starch, biotic transformations may cause significant changes in their behaviour during the exposure period. In general, the two-way interactions between nanoparticles and organisms (potential toxic effects versus particle transformations) constitute an area that is not yet well understood or described in the scientific literature.

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4. Nanoparticles in chemical mixtures and the consequences for bioaccumulation and toxicity As it was described in Chapter 3, the extent and mechanisms of toxicity vary greatly between different types of nanoparticles and between test species and current test procedures are not appropriate to addressing adequately all aspects of the biological effects of these particles. Aside from investigating the ecotoxicological effects of nanoparticles as single compounds, their interactions with co-existing chemicals are also important when evaluating their potential environmental effects and further add to the complexity. Lessons learned from past experience with chemical mixtures have shown that even chemicals of relatively low toxicity towards aquatic organisms as single substances may cause severe effects in chemical mixtures (Kortenkamp et al., 2009). As for conventional water-soluble chemicals, nanoparticles will form part of a complex mixture in the aquatic environment, rather than existing as single contaminants. In most risk assessment guidelines, such as those issued in connection with the two major pieces of European legislation for upstream and downstream chemical regulation, REACH and the Water Framework Directive (WFD), risk assessments are primarily based on the effects of single substances. This approach to risk assessment is considered to be too simplistic, such that it could result in underestimation of the impacts on humans and the environment of chemical compounds (Syberg et al., 2009; Kortenkamp et al., 2009) and likewise underestimate the effects of nanoparticles. The occurrence of nanoparticles in chemical mixtures may originate from the particle production stage, such as particle synthesis and functionalisation. Through their use in, for example, consumer products, cosmetics, medical and remediation applications, nanoparticles will come into contact with other chemical substances such as preservatives, surfactants and active ingredients in pharmaceutical drugs. Finally, through different disposal routes, nanoparticles will come into contact with environmental contaminants present in, for example, wastewater streams and landfill leachate. As a consequence, both intentional and unintentional mixing of nanoparticles with other chemical compounds takes place before, during and after their intended use. Finally, the manufacturing of doped or coated nanoparticles can also be seen as the production of intentional mixtures, whereby an inert and low-toxicity particle could potentially become a carrier of toxic compounds (Hartmann & Baun, 2010 – Paper II). Chemical interactions and mixture toxicity is an area of extensive research where particular attention has thus far been focussed on pesticides, heavy metals,

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endocrine disrupters, poly-aromatic hydrocarbons and general industrial chemicals. Significant scientific progress within this field during the last decade has made possible analysis of mixture effects in multi-component mixtures (Kortenkamp et al., 2009). The presence of nanoparticles in chemical mixtures represents a new challenge since it introduces a solid particle phase, which makes them different from mixture studies involving the aforementioned chemical types. For water-soluble chemicals, a combined mixture effect does not necessarily require a direct physical interaction between the individual compounds. This is the case when the presence of one compound promotes the uptake of another, resulting in synergistic effects ‒ or when compounds are competing for the same binding site, resulting in less than additive effects. In order to describe the role and significance of nanoparticles in mixture interactions, additional scenarios must be included as a result of the presence of a solid particle phase. The relevance of dealing with nanoparticles in chemical mixtures is justified by their specific properties: A large surface area relative to mass which makes them favourable as sorbents and a small size which may allow uptake into organisms, organs and cells (Hartmann & Baun, 2010 – Paper II; Zhang et al., 2007; Zhang et al., 1999). The non-selective uptake of some nanoparticles into the gut of sediment and filter feeders, as well as attachment to algal cell surfaces, further underlines its significance. Figure 7 illustrates conceptually possible interaction scenarios between nanoparticles, environmental contaminants and aquatic organisms. These scenarios are divided into four groups covering four types of interaction, namely a) no interaction, b) interactions increasing overall bioavailability/toxicity, c) interactions reducing overall bioavailability/toxicity and d) interactions which can either lead to increased or reduced overall bioavailability/toxicity. Examples of experimental findings that underpin most of these conceptual interaction scenarios have been described in Hartmann & Baun (2010 – Paper II) focusing on the changes in effect of soluble chemical compounds as a result of the presence of nanoparticles. It should, however, be added that the effect of the chemical compound may also facilitate uptake of nanoparticles as a result of, for example, membrane damage. Hence changes in bioavailability go both ways. In addition, for some of the scenarios, the effect of nanoparticles on, for example, nutrients may be described in a comparable way to that for co-contaminants. This is the case for example if nutrient availability is reduced by the presence of nanoparticles. Finally, the body of literature in this special field is steadily growing, providing additional tangible empirical examples of the conceptual scenarios. In nanoparticle mixture studies, TiO2 is so far the predominant

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nanoparticle type and especially its influence on the bioavailability of heavy metals has been investigated.

Figure 7. Different scenarios for nanoparticles (NP) acting as modifying factors on the effect of a co-contaminant (CC) towards (mainly unicellular) organisms. Scenario 1: Sorption of the CC onto NP. The bioavailability (and effect) of the EC is reduced. Scenario 2: No interaction: effect of the CC is unchanged and NP has no effect on the organism. Scenario 3: No interaction. Effect/uptake of both independently. Scenario 4: Interaction. Increased bioavailability of CC due to increased local CC concentration (or at least adsorbed CC is bioavailable). Scenario 5: Interaction. Increased bioavailability as a result of membrane rupture (caused directly or indirectly by NP or CC) leading to increased uptake and/or effect of CC and/or NP. Scenario 6: Interaction. EC uptake facilitated by NP (Trojan Horse) and increased CC body burden. Scenario 7: Release of free ions from NP, leading to competition with CC (in this case a metal) for binding sites. Reduced CC uptake and effect. Overall effect on the organism may be unchanged, reduced or decreased depending on the specific CC and NP. Possible inherent effect of the NP on the organism. Scenario 8: NP transforming CC to a different product or changing speciation, which results in a product or a species that can either be more or less toxic to the organism. Modified from Hartmann & Baun. (2010 – Paper II). The type of interaction between nanoparticles and co-contaminant will be governed by their individual physicochemical properties. These properties may in some cases limit interactions (Scenarios 2 & 3). This was found to be the case for atrazine and methyl parathion, where limited sorption to nC60 (~10%) and no statistically significant changes in toxicity to D. magna and P. subcapitata were observed. In the same study, 85% sorption of phenanthrene to nC60 was found,

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accompanied by increased toxicity to P. subcapitata. The toxicity to D. magna was also increased 10-fold when concentrations were expressed as water-phase concentrations instead of total concentration. This indicates that phenanthrene sorbed to nC60 aggregates was bioavailable to the organisms (Scenario 4) (Baun et al., 2008b – Paper VII). In another study, the effect of cadmium on the growth rate of P. subcapitata was investigated both in the absence and presence of 2 mg/L TiO2 using three different primary particle sizes (<10, 30 and 300 nm). It was found that the presence of TiO2 resulted in a decreased bioavailability and hence toxicity, of cadmium due to sorption of Cd2+ onto TiO2 (Scenario 1). However, for the 30 nm nanoparticles, growth inhibition was more pronounced than could be explained by the concentration of dissolved Cd(II) species (determined by geochemical modelling), indicating a possible combined effect. This was hypothesised to be either due to bioavailability of cadmium adsorbed to TiO2, increased local concentrations resulting from nanoparticle adhesion to algal cells (Scenario 4), an actual carrier effect (Scenario 6) or combined toxic effect (such as increased cadmium uptake due to cell membrane damage caused by TiO2) (Scenario 5). Conversely, the addition of larger (300 nm) TiO2 particles was found to reduce toxicity of cadmium through sorption resulting in reduced bioavailability (Hartmann et al., 2010 – Paper I) (Scenario 1). Additionally, it has been shown that, despite significant adsorption of cadmium onto TiO2 nanoparticles (>26%), neither accumulation, depuration, acute toxicity nor total body burden in L. variegatus was affected after 24 hours exposure by the presence of TiO2 nanoparticles, compared to cadmium-only exposure. This indicates that the adsorbed cadmium is bioavailable (Scenario 4) to the same degree as in the absence of nanoparticles. In D. magna the total body burden was increased (~ x5) in the presence of TiO2 nanoparticles (Scenario 6) but still no changes in toxicity were seen (Figure 8) (Hartmann et al., 2011c – Paper V).

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Figure 8. Mass balance for uptake of cadmium after 24 and 48 hours exposure to 100 μg/L cadmium in Lumbriculus variegatus (A) and uptake of cadmium after 24 hours exposure to 100 μg/L in Daphnia magna (B) in the absence and presence of 2 mg/L TiO2. Also shown are μXFR images of elemental distribution of Ti (white/light grey) in L. variegatus (head end) (C) and D.magna (D). Together with Ti, the distribution of Na is shown for D.magna (grey) and Si is shown for L. variegatus (grey), outlining the shape of the organisms (both images are ~1x1 mm). (Modified from Hartmann et al., 2011c – Paper V) The function of nanoparticles as carriers for co-contaminants in fish has been demonstrated by Zhang et al. (2007), Sun et al., (2006) and Sun et al. (2009), where the presence of TiO2 nanoparticles was found to increase the accumulation of cadmium and arsenate in carp (Cyprinus carpio) (Scenario 6). TiO2 nanoparticles were found to have a stronger sorption capacity for cadmium than natural soil particles, resulting in a greater accumulation compared to that seen in the presence of natural sediment particles. Besides uptake into the gut, skin and scales – which could be partly reversible ‒ increase in muscle bioconcentration of cadmium was also seen, indicating actual uptake (Zhang et al., 2007). Though

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TiO2 nanoparticles might not themselves translocate into other parts of the animal beside the gut, the co-contaminant can be released inside the gut and move to other organs, resulting in increased bioaccumulation (Sun et al., 2009). Conversely, it was found that bioavailability of 17 alpha-ethinylestradiol, introduced to D. rerio through dietary exposure, appeared to be reduced ‒ or even eliminated ‒ due to association with nC60 aggregates (Scenario 1) (Park et al., 2010). Though no empirical examples have been described in the scientific literature to date, the release of metal ions from nanoparticles (such as Ag and ZnO) could result in competition with other metal ions in chemical mixtures. The presence of metal or metal oxide nanoparticles releasing metal ions can thereby reduce the bioavailability and toxicity of other metals (Scenario 7). The overall change in toxicity will depend on the toxic effect of the two individual metal ions. This may further be a toxic mechanism as release of metal ions from nanoparticles can interfere with nutrient uptake. The toxicity of Ag nanoparticles has thus been linked to the release of Ag+ ions (Miao et al., 2009; Navarro et al., 2008a) which, in turn, influences sodium transport (Li et al., 2010; Bianchini et al., 2002). Finally, some nanoparticles may cause a transformation of a co-contaminant into a different product or species, which then might be either more or less toxic to the organism (Scenario 8). It has recently been found that the presence MWCNTs in non-toxic concentrations increased uptake and toxicity of copper. The increased mortality was attributed to a decreased binding of Cu to NOM. In the presence of MWCNT and NOM the speciation of Cu was changed and an increased concentration of Cu2+ was measured compared to when only NOM was present (Kim et al., 2009). Conversely, the presence of TiO2 modified with thiophene oligomers reduced the inhibitory effect of pentachlorophenol on algal growth by photocatalytic decomposition of pentachlorophenol to less toxic products (Ševčík et al., 2009). These results underline the fact that, in addition to particle toxicity, potential interactions with existing environmental contaminants are also important in assessing the potential environmental risks of nanoparticles. From the point of view of an organism, interactions between nanoparticles and co-contaminants can be beneficial as well as negative, but are in both cases difficult to predict given our present state of knowledge. As mentioned in Section 3.5, a high degree of uptake takes place into the gut of some organisms. Also, as described in Section 3.2., some types of nanoparticle have been found to disturb cell membranes and

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cell walls. Furthermore, as outlined in this section, a high sorption capacity for, for example, heavy metals and organic chemicals has also been observed. In combination these three factors herald a warning as nanoparticles become possible carriers of co-contaminants into organisms potentially resulting in translocation and cellular uptake. In addition to the fact that nanoparticles may influence bioavailability of co-contaminates, the effects of the co-contaminant may also facilitate, for example, cellular uptake of nanoparticles. Binary multi-dose interaction studies may represent a future perspective and one study has in fact already been published by Kim et al. (2010). However, our current level of understanding with regard to nanoparticle behaviour and biological interactions in aqueous test systems is most probably insufficient to ensure that results of such tests can be thoroughly understood. With increased knowledge concerning appropriate test methodology procedures the role and consequence of nanoparticles in chemical mixtures will also become better understood.

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5. Test methods for hazard identification –current applicability and future challenges The basis for discussing and evaluating the potential negative environmental effects of engineered nanoparticles is the ability to test for their hazard potential, including their ecotoxicity and bioaccumulation. The current test paradigm for environmental risk assessment is largely based on the assumption that the chemical substances are water soluble; this will not generally be the case for nanoparticles. As discussed in chapter 3, a number of ecotoxicological studies have demonstrated that it is possible to obtain dose-response relationships when exposing freshwater organisms to aqueous suspensions of nanoparticles. Similarly, bioaccumulation tests have been performed resulting in the calculation of BCF values (Zhu et al., 2010; Tervonen et al., 2010). However, a great degree of variation is observed in toxicity test results, even for tests carried out under comparable test conditions with the same particle types and test species (Menard et al., 2011). As a result of the very different nature and behaviour of nanoparticles compared to soluble chemicals – something which will be discussed further in the next section – the main issue is, therefore, the extent to which such data provide meaningful information on the ecotoxicity and bioaccumulation of nanoparticles and to which results are hampered by unsolved technical and scientific challenges in hazard identification procedures that must be identified and overcome. 5.1. Behaviour of nanoparticles in aquatic test systems The pursuit of an understanding of nanoparticle behaviour in water is prompted by two different information requirements. Firstly, the need to describe and control nanoparticle behaviour in standard tests systems and, secondly, to evaluate their fate in aquatic environments, including identification of the controlling factors. The former is important for obtaining meaningful information on environmental effects and the latter will enable more accurate estimates of environmental concentrations in the different environmental compartments. Together, these values will also feed into the environmental risk assessment of nanoparticles. Nanoparticles in aqueous media are, by nature, different from chemical solutions, resulting in differences in behaviour and interactions with living organisms. In aqueous media, nanoparticles can form colloidal dispersions or suspensions. Colloidal dispersions comprise small particles which are stable in a liquid and cannot be detected under normal light conditions but are large enough to scatter a

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beam of intense light (Zumdahl, 1998; Atkins, 1990). Suspension is a wider term, and suspensions can also consist of larger-sized particles with varying degrees of stability and rate of sedimentation (Zumdahl, 1998). Some types of nanoparticle are typically produced as stable colloidal dispersions by the addition of stabilising agents in particle synthesis (e.g. Ag and Au nanoparticles). Others are industrially produced as dry powders. Nanoparticles added as a dry powder to an aqueous phase will, in many cases, form suspensions of solid particles, in which the stability (aggregation and sedimentation) will be influenced by both particle and media characteristics (von der Kammer et al., 2010; Gao et al., 2009). Examples of the varying potential behaviour of nanoparticles in aquatic media are presented in Figure 9. The transition from solution to colloidal dispersion and to nanoparticle suspension, based on size definitions, is somewhat gradual and blurred, partly due to a lack of consensus. For example, depending on the source of information, colloidal particles are defined as organic or inorganic microparticles or macromolecules with a lower size boundary of one to a few nanometres and an upper size limit of 0.5 to a few micrometres (Hochella, 2008; Oberdörster et al., 2007; Stumm & Morgan, 1996; Schwarzenbach et al., 1993; Atkins, 1990). The dissolved fraction has often been defined operationally by its ability to pass through a filter of a specific pore size, e.g. 0.4-0.45μm (Nowack & Bucheli, 2007; United States Environmental Protection Agency, 2003; Mackay & Fraser, 2000) and hence overlaps with the size range of colloids. For tests with nanoparticles some studies have instead applied centrifugal filtration through 1-2 nm membranes to avoid passage of fine-grained particulates (e.g. Poynton et al., 2011; Miao et al., 2010). A solution has also simply been defined as a homogeneous mixture (Zumdahl, 1998). From a molecular point of view, however, dissolution has been defined as when “individual molecules of the solute are separated by the molecules of the solvent”. This further implies that “the solvent molecules must form a set of bonds with the solute molecules which are, in total, stronger than that of the solute-to-solute bonds. If this does not occur, the solute molecules will move together due to mutual attraction and the substance will come out of solution and thus be insoluble” (Connell, 1997). From this definition of solubility it is clear that nanoparticles are different from soluble chemicals since they are in fact solid-state compounds (e.g. crystalline solids, metals, metal alloys). Fullerenes represent a special case as they are defined as molecular solids, which means that they are held together by strong covalent bonds within the molecule but have quite weak forces between molecules (Zumdahl, 1998). The main fundamental difference between water-soluble

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chemical compounds and nanoparticles is, therefore, the fact that nanoparticles represent a solid phase with a confined physical shape. This results in a boundary between a solid and a liquid phase, which can be seen as a system dominated by interfacial phenomena. Such a system is more concerned with physical forces than molecular transformations (Žutić & Svetličić, 2000), making it distinctly different from chemical solutions. Intermediates are seen in the types of nanoparticles which dissolve partly and release ions.

Figure 9. Different behaviour of nanoparticles in aquatic systems. Particles may form either suspensions of aggregates (A) or agglomerates (B), colloidal suspensions (sols) (C), partly dissolve (D) or dissolve (E). Examples of different nanoparticle behaviour in water are uncoated TiO2 nanoparticles (A/B), coated gold nanoparticles (C), ZnO nanoparticles (D) and nanoparticles of, for example, soluble salts (E). The interactions between two interfaces in a dispersed system, such as two nanoparticles, can be described by the DVLO theorem (Chen & Elimelech, 2007; Feiler et al., 2000). In brief, this says that the overall interaction energy between two interfaces is the sum of attractive van der Waals forces and the repulsive electrostatic Coulomb (double layer interaction) forces. The overall force determines the distance between the particles and whether they form stable dispersions, agglomerate or form aggregates in aquatic media (Salager, 1994). The function of stabilising agents in nanoparticle suspensions or colloidal dispersions is to overcome the attractive forces by additional steric repulsion. Stabilisation can also be achieved by increasing electrostatic repulsion, for example by adjusting the pH to well above or below the isoelectric point (at which total positive charges equal total negative charges) (Hang et al., 2009). Due to the different nature of nanoparticles compared to water-soluble chemicals, they behave differently in aqueous media. Examples of this are shown in Figure 10. It should, however, be kept in mind that there are borderline cases (e.g. partly soluble nanoparticles) and that the differences described here are examples of some general distinctive differences. As described above, the behaviour of water-soluble chemical compounds (including metal compounds) is governed mainly

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by molecular processes such as dissolution, speciation and dissociation as well as sorption (as sorbant). Nanoparticles are different from water-soluble chemicals as they disperse rather than dissolve and as their behaviour is governed by physical forces and interfacial phenomena.

Figure 10. Examples of major general differences between processes of water soluble chemicals and inert nanoparticles in aquous media. Since these forces depend on both particle and media characteristics, media composition will be a controlling factor in particle behaviour. Also the likelihood of particles being in contact with other particles (particle collision) influences aggregation behaviour. Hence nanoparticle concentration is another important factor governing nanoparticle behaviour (Tiede et al., 2009). This has been demonstrated by Hartmann et al. (2011b – Paper IV) where increased sedimentation of TiO2 nanoparticles in OECD algal test media was seen for increasing concentrations (10-100 mg/L). At the same time, suspensions of 100 mg/L TiO2 were relatively stable in MilliQ water over a period of 10 hours, compared to the ionic algal media (Figure 11). Also the sedimentation rate of metal oxide nanoparticles (TiO2, CeO2 and ZnO) has been found to depend on the initial concentration of nanoparticles, with higher stability in natural fresh water compared to sea water (Bai et al., 2010; Keller et al., 2010). Increased sedimentation rate is indirectly a result of increased aggregate size. This is therefore comparable to results obtained by Bai et al. (2010), who observed a 20-fold increase in the hydrodynamic diameter of ZnO nanoparticles (37 nm) when the concentration was increased from 5 to 100 mg/L. Also in a study by Johnston et al. (2010) it was found that aggregation of CeO2, TiO2 and ZnO increased over time and was influenced both by concentration and media type.

Water-soluble chemical compounds

Molecular processesDissolution

Complexation / DissociationDiffusion

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Nanoparticles

Interface processesDispersion

In-situ functionalisation(Dis-)stabilisation

(Dis-)aggregation/agglomerationSedimentation

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Figure 11. Reduction of absorbance (λ=338 nm (TiO2) and λ=523 nm (Au)) as a result of sedimentation of Au and TiO2 nanoparticles suspended in different media and concentrations. Sample 1 is absorbance of a colloidal dispersion of Au nanoparticles diluted to 10 mg/L in algal medium. Samples 7 and 8 correspond to 10 and 100 mg/L TiO2, respectively, suspended in MilliQ water. Samples 2, 3 and 6 all corespond to 10 mg/L TiO2 suspended in OECD algal test medium. Samples 4 and 5 are 40 and 100 mg/L TiO2, respectively, in OECD test media. (Hartmann et al., 2011b – Paper IV). The influence of pH and ionic strength on the aggregation kinetics of nanoparticles has recently been reviewed by Lin et al. (2010) and investigated experimentally for TiO2 nanoparticles by Chen & Elimelech (2007), French et al. (2009) and von der Kammer et al. (2010). In one study it was found that the stability of C60 nanoparticles became increasingly stable at higher pH due to increased negative zeta potential (i.e. surface charge) (Chen & Elimelech, 2007). It has also been demonstrated that TiO2 nanoparticles rapidly form micron-sized aggregates in aqueous suspensions with very low ionic strength in pH ranges of 5.8-8.2, which did not dis-aggregate upon sonication. At pH 4.8, which is below the isoelectric point, TiO2 nanoparticles formed small stable aggregates. Increasing the ionic strength at fixed pH 4.8 led to formation of aggregates within 5 or 15 minutes, with a faster rate being obtained by addition of CaCl2 cation compared to NaCl (French et al., 2009). After measuring TiO2 nanoparticle stability, zeta potential and aggregate size as a function of variations in pH and water composition it was proposed by von der Kammer et al. (2010) that particles of the same core material will behave in a similar predictive way. This study further highlights the complex interactions between water chemistry parameters and nanoparticle behaviour, as well as finding that NOM was a

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stabilising factor under all the tested conditions of varying pH and concentrations of mono- and divalent cations. It can be argued that ‘stable colloidal suspension’ might be the nanoparticle equivalent of solution. Nonetheless, there are major differences between the behaviour of conventional chemicals and nanoparticles in aqueous media (Figures 9 & 10) and their interactions with biological systems (Figures 3 & 4). When testing soluble nanoparticles and nanoparticles releasing metal ions we can – at least to some extent – build on our experience with conventional soluble organic chemicals and metals. However, the introduction of particle suspensions in commonly used aquatic test systems presents a major challenge and many questions remain to be answered. 5.2. Exposure quantification in a dynamic system Aggregation of nanoparticles is a general issue that is observed and described in the majority of published studies. Even in studies focused on one specific particle type, differences in media composition and suspension preparation procedures introduce significant variability into nanoparticle characteristics and behaviour, which is likely to influence test results. Time- and media-dependent aggregation – and even non-linear concentration-aggregation relationships – is a problem that has to be overcome in order to obtain reliable and comparable test results (Baun et al., 2009). Monitoring changes in water-phase concentrations is an obvious option but is hampered by grazing behaviour (e.g. for daphnids) and adhesion and co-sedimentation (e.g for algae), which will result in under-estimation of exposure. Measuring the water concentration of nanoparticles during the course of the test will clearly give useful information on nanoparticle stability but not necessarily express actual exposure levels. Due to the dynamic nature of the test systems and changes in concentrations resulting from, for example, sedimentation and biomodification (as discussed in Sections 3.6. and 5.1.), organism burden could be an alternative to water-phase concentrations for some organisms. Some nanoparticles (such as metal oxide nanoparticles) have a tendency to form aggregates and combine with algal cells. As a result of this and the possible role of direct physical cell-particle contact in effect mechanisms, degree of adhesion could be used as an indicator for the potential of nanoparticles to interact with cells and be a way of quantifying algal exposure. Quantification of the degree of adhesion could possibly be acheived by immobilising algae on a support matrix, allowing their separation from media and suspended nanoparticles. Since the biological interactions of nanoparticles will vary as a result of many, some as yet unidentified, parameters, such methods

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may provide a way of performing an initial screening, thereby determining the degree of direct interaction which is likely to be a prerequisite for some types of particle-mediated effects on algae. A somewhat similar concept has been proposed by Xia et al. (2010) via a biological surface adsorption index (BSAI) approach. By quantifying the adsorption coefficients of small molecules and biomolecules, of known characteristics, on a given nanomaterial, so-called BSAI nanodescriptors can be established. It is proposed to use this as a tool to predict membrane interaction and biodistribution parameters, including cellular uptake of nanoparticles (Xia et al., 2010). Based on published studies, simply expressing nanoparticle exposure to filter and sediment feeding organisms as a mass water phase concentration does not seem to be entirely appropriate either. This conclusion is partly based on the fact that particle properties (and hereby effects) are influenced by the organism itself, for example by excretion of organic matter. Also the large uptake into the gut observed with respect to some nanoparticles raises questions concerning whether it is internal or external exposure that will predominantly determine effects. For filter and sediment feeding organisms their gut content of nanoparticles will often serve as a large intra-organism source of nanoparticle exposure. Hence correlating total body burden to effects or translocation may provide more meaningful results than qualitative and quantitative water-phase exposure measurements. If correlations between body burden and toxicity can be found, this places additional emphasis on finding the controlling factors behind particle uptake and their translocation (as discussed in Section 3.5) to other organs as these two factors will then, in combination, be the cause of the observed adverse effects. In studies of particle uptake, not only concentration, but also exposure volume has been found to influence the degree to which this occurs. Increasing test suspension volumes (and hereby total CNT mass) caused significantly increased uptake of CNTs in D. magna (Petersen et al., 2009). Another possible way of addressing the issue could be to test a surface of nanoparticles. This would, however, prevent some types of effect, such as those related to internalisation and translocation. This procedure may, however, provide information regarding the effects of, for example, nano-structured surfaces and ion release. Finally, there is the problem of concentration-dependent aggregation as described by, for example, Baloousha et al. (2009), Hartmann et al. (2010 – Paper I), Hartmann et al. (2011b – Paper IV) and Bai et al. (2010), which means that exposure also changes qualitatively as a result of quantitative increase in

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concentration. The standard approach to concentration response testing does, hereby, include interdependent variables (e.g. concentration and aggregation state). As a result, inhibition found in the test cannot be confidently claimed to be solely dependent on nanoparticle concentration (Hartmann et al., 2010 – Paper I). In other cases, nanoparticles may also change in a time-dependent manner during the course of the test – not only in aggregation state but also in composition. This might be the case for coated nanoparticles. For example stable colloidal dispersions of starch-coated nanoparticles have been found to decrease in size during exposure – which is probably related to a decrease in coating thickness – and form aggregates after 48 hours of incubation, which again is thought to be related to degradation of the starch coating (Hartmann et al., 2011b – Paper IV). The fact that exposure will, in some cases, change both quantitatively and qualitatively as a function of time and concentration, may well require alternative approaches to be adopted in order to understand links between particle characteristics and the resulting effects. Hence, both quantitative and qualitative descriptors of exposure should be monitored and described due to the dynamic nature of the test system. Measurements of nanoparticles, as well as test system parameters, should be carried out regularly throughout the test period. Some inspirational guidance regarding monitoring of test concentrations during exposure is provided by the OECD guidance document on poorly soluble substances (OECD, 2000). Other measures for nanoparticles exposure, aside from water-phase concentrations, should be investigated and could be based on actual organism burden. 5.3. Correlating particle characteristics to effects Effects of nanoparticles on aquatic organisms, as described in the scientific literature, include both those believed to be directly related to the presence of the nanoparticles and their specific properties, but also those related to, for example, ion release, the presence of impurities or solvents as well as physical hindrance. An understanding of the effects of well-known chemical components resulting from nanoparticle exposure, such as released metal ions, is supported by a long history of scientific investigation. However, additional particle-specific effects, resulting from, for example, interfacial phenomena, are not well understood. The interface interactions between nanoparticle and biological surfaces have been explored conceptually by Nel et al. (2009) and – as it was described for inter-particle interactions in Section 5.1. – comprise biophysicochemical interactions at the interface between the two. Understanding the interactions and effect mechanisms is also a step towards choosing suitable dose descriptors, though possibly complicated by a combination of several intrinsic properties influencing

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biological effects. This also makes it difficult to find common dose descriptors for nanoparticles in general since they have very different ways of affecting aquatic organisms. For some, ion release is relevant, for others morphology or crystallinity might instead be a determining factor. Some studies have found surface area to be an appropriate dose descriptor to explain size dependant changes in SiO2 toxicity to algae (van Hoecke et al., 2008) as well as CeO2 toxicity to algae and chronic toxicity to D. magna (van Hoecke et al., 2009). It was shown that effects increased with decreasing particle size in a way that could be explained by an increase in surface area. Hence, in this case, surface area might be a more suitable measure of exposure compared to concentration usually given in mass per volume (Baun et al., 2009). The significance of surface area for toxicological effects has clearly been demonstrated in inhalation studies where increased surface area has been linked to, for example, increased fibrogenicity (Oberdörster et al., 2007). However, such correlations have not generally been described in ecotoxiclogical tests although surface area is without doubt one of the decisive parameters for particle reactivity and, hereby, potentially also effects. A lower degree of aggregation may result in more direct links between surface area and effect. Correlation of effects to aggregate sizes has been proposed by Menard et al. (2011), although this was hampered by current data availability. However, as a decrease in particle size has been found to increase aggregation (Penn et al., 2007), whereas increased surface charge has the opposite effect (Hang et al., 2009), this approach may not be so straightforward. Furthermore, there can be differences in compactability resulting from variations in particle size and shape. Smaller particles may, for example, aggregate in a more compact way than larger ones. There may also be differences in surface roughness which is believed to influence interactions at the bio-nano interface (Nel et al., 2009; Yeung et al., 2009). Though aggregates of smaller nanoparticles may still exhibit a larger degree of adverse effect towards organisms compared to aggregates of larger particles, the increased effect will probably not be proportional to the smaller size of the primary particles. Again, this approach is also hampered by concentration-dependent aggregation which makes it difficult to establish dose-effect relationships. The presence of NOM has also been found to alter the toxicity of both metals and metal oxides (Miao et al., 2009; Hall et al., 2009; Koukal et al., 2007). Accordingly, since the properties of the medium influence particle properties and metal speciation and hereby interface interactions and toxicity, relating effects directly to nanoparticles properties is further hampered.

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By separation of different size fractions of nanoparticle suspensions it has been found that smaller size fractions seem, in some cases, to be responsible for the observed effects (Hund-Rinke et al., 2010). Hence, for nanoparticles forming micron-sized aggregates, the overall effect may be predominated by a small fraction of the nanoparticles present as single particles or small aggregates. To determine if this is a general trend, the testing of separated size fractions may aid a better understanding of the toxic mechanisms. The fact that overall effects can be caused by two or more separate contributions is also pointed out by Miao et al. (2009). In this case, however, the point is that direct nano-specific effects might be ‘camouflaged’ by other effects such as ion release. In general, it is a challenge to distinguish indirect effects caused by changes to the cell environment from direct effects caused by direct physical or chemical interactions at the interface between nanoparticle and cell. Bringing together some of the issues highlighted in this section, efforts should be concentrated on understanding the interactions between nanoparticles and organisms, both at a cellular level (as described by Nel et al. (2009)) and also at an organism level. Due to the differences in nanoparticle properties as well as behaviour and effect mechanisms, a fundamental understanding of the properties governing these factors is needed. Once this is better understood, the choice of dose descriptor might, in some instances, be performed on a case-by-case basis for different groups of nanoparticles with similar modes of effect. Reporting the dose, expressed on a weight basis, supplemented with sufficient characterisation data, enables re-calculation into various alternative units (OECD, 2010b). As pointed out by several authors, the reporting of nanoparticle effects without thorough characterisation is, in all cases, of very limited value for data interpretation and future data evaluation (e.g. Menard et al., 2011; Tiede et al., 2009; Watheit et al., 2007). Characterisation should not be limited to initial measurements but should instead be performed continuously throughout the test period. The numerous aspects related to nanoparticle behaviour, exposure characterisation and possible links to toxic effects, highlight the highly inter-disciplinary nature and complexity of this research field. As a consequence, research projects require cooperation between scientists with backgrounds in biology, chemistry and colloidal science. Adoption of such a wide interdisciplinary approach is crucial to the gaining of an understanding of the behaviour and effects of nanoparticles in aquatic test systems, which is necessary

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to solve the major challenge of controlling and describing both dose and effects in a meaningful way. 5.4. Adaptation of current testing methods – what is the potential? The current variability in results of ecotoxicological tests relating to nanoparticles can be illustrated, to some extent, by TiO2 nanoparticles, where 48 hour EC50 values for acute toxicity to D. magna for TiO2 nanoparticles (≤20 nm – 70 nm, suspensions prepared in a comparable manner by sonication) have been reported to vary between 35 and 20.000 mg/L) (Zhu et al., 2010; Zhu et al., 2009; Heinlaan et al., 2008; Lover & Klaper, 2006). By increasing the test period from 48 to 72 hours, Zhu et al. (2010) observed a decrease in EC50 value from >100 mg/L to 1.62 mg/L. The issue of low reproducibility in algal growth inhibition tests has also been pointed out by Hartmann et al. (2010 – Paper I). It has further been concluded by Menard et al. (2011) that currently available data on TiO2 nanoparticle toxicity to P. subcaitata are, in general, extremely variable. Differences in particle characteristics can form part of the explanation, although they are unlikely to be the sole cause. Although these variations in effect for seemingly comparable tests are more extreme than for other nanoparticle types, it still points to the fact that commonly used test procedures may not be sufficiently robust to deal with nanoparticles. Consequently, common procedures for, for example, suspension preparation methods and exposure conditions are required. These factors could then, at least, be ruled out when comparing results between tests. Inaptness of the test systems is also related to the fact that exposure conditions change qualitatively as a function of time and concentration for the same nanoparticle type. A fundamental question regarding dynamic exposure conditions is also whether the aim is to describe or to control. A way of controlling exposure is through stabilising the test systems by different means, although this might not always prove to be an appropriate solution. As mentioned above, the stabilisation of nanoparticles with NOM has been found to reduce metal toxicity (Miao et al., 2009; Hall et al., 2009; Koukal et al., 2007). Consequently, it may not be appropriate for the testing of ion-releasing nanoparticles. Media renewal is also an option. However, due to the large degree of uptake into the gut by organisms and adhesion to biological surfaces, this may increase body-burden and thereby internal exposure. This factor is also something which should be kept in mind in long-term chronic tests. Alternative test strategies to control exposure may

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include the use of short-term tests. This would help to minimise the influence of time-dependent changes in exposure conditions and the effects that the organisms themselves have on nanoparticle exposure. However, it requires more sensitive methods for the monitoring of short-term sub-lethal or sub-inhibitory biological effects able to detect small changes in response. A draw-back of short exposures is that they may be insufficient for investigating the potential adverse effects of pristine nanoparticles. As described by Tiede et al. (2009), the situation is analogous to that seen in pesticides, where the fate and effect of the altered compound should also be examined as potentially harmful effects may otherwise be overlooked. Another option is to test at single concentrations – or possibly several low concentrations within a limited interval. The aim here is again to ensure as homogenous qualitative exposure conditions as possible. However, this would allow mainly ‘between-particle-type’ comparison and conflicts with the concept of dose-effect testing. Complexity of the test system can also be reduced by isolating specific size fractions of nanoparticle aggregates. For instance, it is recommended by ISO (2006) that effects of poorly soluble inorganic materials should be determined by testing water-soluble fractions (WSF) prepared by stirring in water and subsequent filtration. Though this approach is relevant for partly soluble nanoparticles, for which dissolved ions may contribute to effects, only testing WSFs is inappropriate for non-soluble nanoparticles where the presence of a particle may be related to its mode of action. Although it has been found by Hund-Rinke et al. (2010) that the smaller fraction of TiO2 particles (passing though a 0.22 um filter) were responsible for toxic effects on P. subcapitata, the reason for this is not well understood. Dilution of the filtrate almost eliminated the effects, possibly as a consequence of aggregation. Based on these results, testing of filtrates is not recommended until effect mechanisms are better understood (Hund-Rinke et al., 2010). However, in combination with testing of unfiltered samples it may provide useful information. If the aim is instead to describe exposure, then the challenge is to find a way of taking the complex changes in particle properties into account. In cases where determining factor relative to the adverse effects has been identified this may be possible by measuring this specific parameter. This approach would then be comparable to the models for metal toxicity (FIAM and BLM) as well as Kow based models for organic chemicals. However, if effects are caused by interacting and non-static parameters this approach would be challenging, to say the least.

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It is possible to draw parallels between the testing of nanoparticles and testing of metals and metal compounds, as described in ISO 14442:2006 (ISO, 2006). Firstly, their behaviour (chemical form and interaction with biological tissue) is controlled by factors such as pH and ionic strength. Secondly, organic matter, such as algal exudate, has been found to be an influencing factor. Another common feature is that the octanol-water partition coefficient is not a valid predictor of the bioaccumulation potential of metals and inorganic metal compounds – and most likely not for most types of nanoparticles, though some Kow estimates have been made for carbon-based nanoparticles (Petersen et al., 2010; Jafvert & Kularni, 2008). However, MWCNT uptake in Eisenia foetida and L. variegatus could not be correlated to the estimated Kow value (Petersen et al., 2010). These compounds are also, by nature, non-biodegradable. For these reasons – and since many nanoparticles are composed of metals and inorganic materials – general guidance on testing such compounds is also highly relevant in a nano-context. Among other things, it is suggested in ISO 14442:2006 (ISO, 2006) that the algal biomass is kept low and that pH is carefully controlled. Minimising changes in pH is crucial relative to controlling metal speciation and with respect to nanoparticles, Similarly, the pH is likely to influence nanoparticle aggregation (Hang et al., 2009) or ion release (Lui & Hurt, 2010). ISO 14442:2006 (ISO, 2006) points out, furthermore, that test results will be specific to the test material and physical-chemical properties of the test medium and this applies similarly to nanoparticles. Algal tests are in many ways different from other aquatic ecotoxicological tests (OECD, 2000) and this also has implications for tests involving nanoparticles. Here, the quantification of biomass is yet another nano-specific challenge as the most common methods employed as a surrogate for biomass based on cell counts (Coulter particle counter and haemocytometer) or pigment extraction (fluorescence) are hampered by nanoparticle background. It can be very difficult to subtract background values due to algae-particle interactions (biotransformation and aggregation formation) and nanoparticle transformation. Based on the findings published in Hartmann et al. (2011b – Paper IV), fluorescence of pigment extracts is suggested as an attractive method as it allows for a physical separation of biomass surrogate (pigment) and particles. It is also proposed that the method can be further adjusted to reduce particle background noise, for example by centrifugation or filtration of the pigment extract. Also, the combination with visual cell counting is recommended as it will help to increase our understanding of particle effects and their interactions with algal cells. An additional endpoint measurement could be the pigment composition as it has

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been used to indicate effects related to irradiance levels (e.g. Dubinsky & Stambler, 2009) including shading of algae at a cellular level by nanoparticle attachments (Wei et al., 2010). Another interesting discussion is related to the choice of test species. It was argued by Baun et al. (2008a – Paper VI) that sediment dwelling organisms may be key test organisms due to the expected environmental fate of nanoparticles. However, also from a test applicability point of view, nanoparticle exposure through sediments will eliminate some of the problems related to keeping the suspensions stable. Still, sediment preparation may be challenging as a homogeneous mixture of sediment and particles can be difficult to obtain. Within one group of organisms, some species may also represent more suitable alternatives than others. For example, algae that are able to carry out endocytosis (such as E. gracilis and O danica) may be more susceptible to nanoparticle exposure and may provide a ‘worst case scanario’ for nanoparticle uptake and the effects caused by this. They may also provide a link between human cytotoxicity and ecotoxicity and render the results more comparable. As has been pointed out throughout the previous chapters, the testing of nanoparticles can currently be described as an equation comprised of many variables and unknowns. The underlying problem is that we are attempting to test hypotheses (the influence of particle characteristics on the effects) and the appropriateness of our test methods simultaneously. One approach is to test for generalising endpoints (such as mortality, reproduction and bioaccumulation) using standardised test methods and well-known test systems for a wide range of nanoparticles and organisms in order to expand the general knowledge base. Such studies can elucidate links between nanoparticle properties and effects but will require careful characterisation of nanoparticle properties, exposure conditions and particle behaviour. This represents a major challenge in a complex and dynamic test system. At the same time, such studies may not necessarily provide much information on the mechanism of (toxic) action, rendering it difficult to clarify the underlying cause-effect relationships. If our aim is to understand the underlying processes, more targeted studies of specific processes (for example by making use of biochemical assays such as the 2′,7′-dichlorofluorescein diacetate ROS assay) are required. Studies of reduced complexity are also needed where the exploration of alternative methods for quantifying nanoparticle effects, such as sensitive short-term assays, may be attractive options. Interdisciplinary research is absolutely key to bringing

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together all the pieces of the puzzle and allowing us to understand the connections between nanoparticle characteristics, behaviour and effects. In the scientific literature these two approaches are increasingly being integrated, as generalising standard endpoints are combined with other (non-standard) endpoints such as gene expression, studies of translocation, cell permeability, biotransformation of nanoparticles, methodological artefacts etc. Furthermore, detailed visual inspection of organisms and particle suspensions as a part of ecotoxicological tests, using methods such as electron microscopy, is more the rule than the exception. This approach, involving taking one step backwards in order to perform an in-depth evaluation of the analytical methods, test procedures and biological processes, is needed to ensure the robustness of the applied test methods and the relevance of test endpoints. In relation to test procedures, some aspects can (and should) be harmonised by additional guidance – especially for use in regulatory testing – to optimise test procedures, ensure comparable results and optimise the quality of the output data. Such guidance should undoubtedly include both quantitative and qualitative exposure descriptors, and a thorough characterisation of nanoparticle properties and behaviour is therefore needed. For example, a stable and homogeneous stock suspension should be achieved and concentrations of stabilising compounds should possibly be constant between exposure suspensions to ensure qualitatively consistent exposure for all tested concentrations. Adaptation of existing test guidelines through the development of nano-specific guidance documents has been proposed by the OECD WPNM as a way of dealing with regulatory ecotoxicological testing of nanoparticles. In the light of the fundamental different nature of nanoparticles, compared to water-soluble chemicals, as well as the suggestions presented in the above section, appropriate testing of nanoparticles in order to identify nano-specific effects may lie beyond the adaptation limits of standard test guidelines. This could result in inadequate protection of human health and environment, which highlights the immediate need for targeted research efforts to increase the knowledge base on which new guidelines or guidance can be developed. The applicability of current test procedures and standard test methods for ecotoxicological testing of nanoparticles is clearly up for scientific and regulatory debate. This is true in relation to future standard test guidelines for use in regulatory testing, as well as to the future evaluation of the European chemical

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legislation on the Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) in 2012, which opens the door for amendments to the legislation in relation to nanoparticles (EurActiv, 2010). A number of important issues have been highlighted within this thesis, which provide input and offer suggestions relative to this ongoing debate.

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6. Conclusion Based on the current body of knowledge arising from (eco)toxicological testing of nanoparticles, there are several reasons why their environmental effects should be of concern. Firstly, large-scale production and persistency of many types of nanoparticles will, over time, lead to their accumulation in the environment. Secondly, although general conclusions on nanoparticle ecotoxicity cannot be drawn due to the extreme diversity of material types and particle properties, some types of nanoparticles have been found to cause acute and sub-lethal effects at low concentrations. As detailed in the preceding chapters, effect mechanisms relating to nanoparticles are still not clearly understood and, consequently, some nano-specific effects might therefore be overlooked through the application of test methods developed mainly for water-soluble chemicals. Understanding the underlying mechanisms relating to the potentially adverse effects of nanoparticles on aquatic organisms is a prerequisite for determining appropriate test strategies. There is, however, still some way to go before achieving this goal and the task will be complicated by the diversity of nanoparticles, their complex nature and behaviour as well as the rapid development of new particle types for an ever-increasing number of applications. This thesis has highlighted and demonstrated the complexity of testing the ecotoxicity of nanoparticles in aquatic test systems. Based on experimental work and the scientific literature it is concluded that the establishment of dose-response relationships is hampered by the dynamic nature of the test systems, including dose-dependent aggregation, as well as by problems in distinguishing direct and indirect contributions to the overall observed effects. Although qualitative links between cause and effect can be established (i.e. nanoparticle exposure causes an effect) this is far from the same as a meaningful dose-response relationship based on aqueous concentrations. This is partly due to the fact that exposure can change qualitatively as a result of changing concentrations, but also to the fact that effects can be mediated through adsorption to cell walls or uptake into the gut. Alternative ways of quantifying exposure should therefore be a topic of future studies. 6.1. Outlook As has been pointed out repeatedly in this thesis, ecotoxicological testing of nanoparticles can currently be best described as an equation comprised of many unknown parameters and variables. In attempting to test simultaneously both hypotheses (e.g. influence of different sizes, coatings, structures etc. on toxicity)

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and the appropriateness of the applied test methods, we are at a great risk of misinterpreting our results. Hence, at our present level of understanding, a narrower focus on the technical, methodological and biological details is required. This includes in-depth examination of, for example, methodological problems, test-system dynamics, ecotoxicokinetics and effect mechanisms. Areas of particular interest in the future could be:

- Biotic and abiotic transformations in test systems during the test period – as nanoparticle suspensions may be highly dynamic.

- Differentiation between the causes of effects (ion versus particle effects; direct versus indirect effects).

- Nanoparticle bioavailability including intra-organism fate and behaviour (ecotoxicokinetics), as well as appropriate methodologies for determining bioconcentration factors for nanoparticles. Standard purging procedures could be an outcome of such work.

- Investigations of alternative dose-metrics, including the use of organism burden to quantify actual exposure. This is relevant due both to the highly dynamic test systems but also the observed uptake into the gut of some test organisms or adhesion of nanoparticles to algal surfaces.

- Biomass – and hereby effect – quantification in algal tests, as these tests represent a special methodological challenge due to problems relating to physical separation of biomass and nanoparticles.

- The particulate nature of nanoparticles should be considered in relation to choice of test species. For example algae capable of endocytosis may be of particular interest.

Finally, the presence of nanoparticles in chemical mixtures has been found to influence bioavailability and thereby toxicity, of the individual compounds. Interaction studies (especially multiple-dose studies) are, in some respects, premature at present as they represent increased (rather than decreased) complexity and engender problems of result interpretation compared to single-compound toxicity testing. However, single-dose interaction studies can provide some indications concerning the potential of nanoparticles to influence mixture toxicity and the mode of their action. Also, when fundamental test procedures are in place, the role of nanoparticles in chemical mixtures will become an important field for future studies.

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References Albanese, A., Sykes, E.A., & Chan, W.C.W. 2010. Rough around the Edges: The inflammatory

response of microglial cells to spiky nanoparticles. ACS Nano 4(5): 2490-2493.

Andersson, A.S., Bäckhed, F., von Euler, A., Richter-Dahlfors, A., Sutherland, D., & Kasemo, B. 2003. Nanoscale features influence epithelial cell morphology and cytokine production. Biomaterials 24(20): 3427-3436.

Andrievsky, G., Shakhnin, D., Tronza, A., Zhernosekov, D., & Tykhomyrov, A. 2010. The acceleration of blood plasma clot lysis in the presence of hydrated C60 fullerene nanostructures in super-small concentration. Fullerenes, Nanotubes and Carbon Nanostructures 18(3): 303-311.

Aruoja, V., Dubourguier, H., Kasemets, K., & Kahru, A. 2009. Toxicity of nanoparticles of CuO, ZnO and TiO2 to microalgae Pseudokirchneriella subcapitata. Science of The Total Environment 407(4): 1461-1468.

Arvizo, R.R., Miranda, O.R., Thompson, M.A., Pabelick, C.M., Bhattacharya, R., Robertson, J.D., Rotello, V.M., Prakash, Y.S., & Mukherjee, P. 2010. Effect of Nanoparticle Surface Charge at the Plasma Membrane and Beyond. Nano Letters 10(7): 2543-2548.

Atkins, P.W. 1990. Physical chemistry. Oxford University Press, Oxford, United Kingdom (ISBN 9780198552840). p.706

Auffan, M., Rose, J., Bottero, J.Y., Lowry, G.V., Jolivet, J.P., & Wiesner, M.R. 2009. Towards a definition of inorganic nanoparticles from an environmental, health and safety perspective. Nature Nanotechnology 4(10): 634-641.

Baalousha, M. 2009. Aggregation and disaggregation of iron oxide nanoparticles: Influence of particle concentration, pH and natural organic matter. Science of The Total Environment 407(6): 2093-2101.

Bai, W., Zhang, Z., Tian, W., He, X., Ma, Y., Zhao, Y., & Chai, Z. 2010. Toxicity of zinc oxide nanoparticles to zebrafish embryo: a physicochemical study of toxicity mechanism. Journal of Nanoparticle Research 12(5): 1645-1654.

Battin, T.J., Kammer, F.V., Weilhartner, A., Ottofuelling, S., & Hofmann, T. 2009. Nanostructured TiO2 : Transport behavior and effects on aquatic microbial communities under environmental conditions. Environmental Science & Technology 43(21): 8098-8104.

Baun, A., Hartmann, N.B., Grieger, K., & Kusk, K.O. 2008a. Ecotoxicity of engineered nanoparticles to aquatic invertebrates: a brief review and recommendations for future toxicity testing. Ecotoxicology 17(5): 387-395.

Baun, A., Hartmann, N.B., Grieger, K.D., & Hansen, S.F. 2009. Setting the limits for engineered nanoparticles in European surface waters – are current approaches appropriate? Journal of Environmental Monitoring 11(10): 1774.

Baun, A., Hartmann, N.B., Hansen, S.F., Peters, S., Hankin, S., Micheletti, C., Sokull-Klüttgen, B., Fernandes, T., Stone, V. 2010. Ecotoxicity of four nanomaterials – the results of the ENRHES review. 2010. In: Nanotoxicology 2010 Edinburgh, Meeting Programme & Abstracts, p. 19. Edinburgh Napier University, Edinburgh.

58

Baun, A., Sørensen, S., Rasmussen, R., Hartmann, N., & Koch, C. 2008b. Toxicity and bioaccumulation of xenobiotic organic compounds in the presence of aqueous suspensions of aggregates of nano-C60. Aquatic Toxicology 86(3): 379-387.

Bernhardt, E.S. et al. 2010. An ecological perspective on nanomaterial impacts in the environment. Journal of Environment Quality 39(6): 1954.

Bhushan, B. 2007. Springer Handbook of Nanotechnology. Springer, Berlin, Germany (ISBN 9783540298557). p.1

Bianchini, A., Grosell, M., Gregory, S.M., & Wood, C.M. 2002. Acute silver toxicity in aquatic animals is a function of sodium uptake rate. Environmental Science & Technology 36(8): 1763-1766.

Bosi, S. 2003. Fullerene derivatives: an attractive tool for biological applications. European Journal of Medicinal Chemistry 38(11-12): 913-923.

Brayner, R., Dahoumane, S.A., Yéprémian, C., Djediat, C., Meyer, M., Couté, A., & Fiévet, F. 2010. ZnO nanoparticles: synthesis, characterization, and ecotoxicological studies. Langmuir: The ACS Journal of Surfaces and Colloids 26(9): 6522-6528.

Brayner, R., Ferrari-Iliou, R., Brivois, N., Djediat, S., Benedetti, M.F., & Fiévet, F. 2006. Toxicological impact studies based on Escherichia coli bacteria in ultrafine ZnO nanoparticles colloidal medium. Nano Letters 6(4): 866-870.

Cañas, J.E., Long, M., Nations, S., Vadan, R., Dai, L., Luo, M., Ambikapathi, R., Lee, E.H., & Olszyk, D. 2008. Effects of functionalized and nonfunctionalized single-walled carbon nanotubes on root elongation of select crop species. Environmental Toxicology and Chemistry 27(9): 1922-1931.

Chen, K.L., & Elimelech, M. 2007. Influence of humic acid on the aggregation kinetics of fullerene (C60) nanoparticles in monovalent and divalent electrolyte solutions. Journal of Colloid and Interface Science 309(1): 126-134.

Cheng, J., Flahaut, E., & Cheng, S.H. 2007. Effect of carbon nanotubes on developing zebrafish (Danio rerio) embryos. Environmental Toxicology and Chemistry 26(4): 708-716.

Connell, D.W. 1997. Basic Concepts of Environmental Chemistry 1st ed. CRC Press / Lewis Publishers, Boca Raton, Florida, United States (ISBN: 978-0873719988). p.17-44

Dubinsky, Z., & Stambler, N. 2009. Photoacclimation processes in phytoplankton: mechanisms, consequences, and applications. Aquatic Microbial Ecology 56: 163-176.

Edgington, A.J., Roberts, A.P., Taylor, L.M., Alloy, M.M., Reppert, J., Rao, A.M., Mao, J., & Klaine, S.J. 2010. The influence of natural organic matter on the toxicity of multiwalled carbon nanotubes. Environmental Toxicology and Chemistry 29(11): 2511-2518.

Elzey, S., & Grassian, V.H. 2010. Agglomeration, isolation and dissolution of commercially manufactured silver nanoparticles in aqueous environments. Journal of Nanoparticle Research 12(5): 1945-1958.

59

Engelbrekt, C., Sørensen, K.H., Zhang, J., Welinder, A.C., Jensen, P.S., & Ulstrup, J. 2009. Green synthesis of gold nanoparticles with starch–glucose and application in bioelectrochemistry. Journal of Materials Chemistry 19(42): 7839.

EurActiv, 2010. REACH register to ensure traceability of nanomaterials. Available at: http://www.euractiv.com/en/food/reach-register-ensure-traceability-nanomaterials-news-497781 [Accessed February 10, 2011].

European Commission. 2008. Commission Staff Working Document: Accompanying Document to the Communication from the Commission to the European Parliament, the Council and the European Economic and Social Committee: Regulatory Aspects of Nanomaterials. Available at: http://eurlex.europa.eu/LexUriServ/LexUriServ.do?uri=COM:2008:0366:FIN:en:PDF [Accessed February 10, 2011].

European Commission. 2010. Review of Cosmetics Directive 76/768/EEC - Cosmetics - Consumer Affairs. Available at: http://ec.europa.eu/consumers/sectors/cosmetics/documents/revision/index_en.htm [Accessed February 10, 2011].

Farré, M., Gajda-Schrantz, K., Kantiani, L., & Barceló, D. 2009. Ecotoxicity and analysis of nanomaterials in the aquatic environment. Analytical and Bioanalytical Chemistry 393(1): 81-95.

Feiler, A., Jenkins, P., & Ralston, J. 2000. Metal oxide surfaces separated by aqueous solutions of linear polyphosphates: DLVO and non-DLVO interaction forces. Physical Chemistry Chemical Physics 2(24): 5678-5683.

Feldheim, D.L. 2007. The new face of catalysis. Science 316(5825): 699 -700.

Franklin, N.M., Rogers, N.J., Apte, S.C., Batley, G.E., Gadd, G.E., & Casey, P.S. 2007. Comparative toxicity of nanoparticulate ZnO, bulk ZnO, and ZnCl2 to a freshwater microalga (Pseudokirchneriella subcapitata): The importance of particle solubility. Environmental Science & Technology 41(24): 8484-8490.

French, R.A., Jacobson, A.R., Kim, B., Isley, S.L., Penn, R.L., & Baveye, P.C. 2009. Influence of ionic strength, pH, and cation valence on aggregation kinetics of titanium dioxide Nanoparticles. Environmental Science & Technology 43(5): 1354-1359.

Gao, J., Youn, S., Hovsepyan, A., Llaneza, V.L., Wang, Y., Bitton, G., & Bonzongo, J.C. 2009. Dispersion and toxicity of selected manufactured nanomaterials in natural river water samples: effects of water chemical composition. Environmental Science & Technology 43(9): 3322-3328.

Gao, L., & Zhang, Q. 2001. Effects of amorphous contents and particle size on the photocatalytic properties of TiO2 nanoparticles. Scripta Materialia 44(8-9): 1195-1198.

Geller, W., & Müller, H. 1981. The filtration apparatus of Cladocera: Filter mesh-sizes and their implications on food selectivity. Oecologia 49(3): 316-321.

http://www.euractiv.com/en/food/reach-register-ensure-traceability-nanomaterials-news-497781�

http://www.euractiv.com/en/food/reach-register-ensure-traceability-nanomaterials-news-497781�

http://eurlex.europa.eu/LexUriServ/LexUriServ.do?uri=COM:2008:0366:FIN:en:PDF�

http://ec.europa.eu/consumers/sectors/cosmetics/documents/revision/index_en.htm�

60

Gillis, P.L., Chow-Fraser, P., Ranville, J.F., Ross, P.E., & Wood, C.M. 2005. Daphnia need to be gut-cleared too: the effect of exposure to and ingestion of metal-contaminated sediment on the gut-clearance patterns of D. magna. Aquatic Toxicology 71(2): 143-154.

Gophen, M., & Geller, W. 1984. Filter mesh size and food particle uptake by Daphnia. Oecologia 64(3): 408-412.

Gottschalk, F., Sonderer, T., Scholz, R.W., & Nowack, B. 2009. Modeled environmental concentrations of engineered nanomaterials (TiO2, ZnO, Ag, CNT, Fullerenes) for different regions. Environmental Science & Technology 43(24): 9216-9222.

Grieger, K.D., Fjordbøge, A., Hartmann, N.B., Eriksson, E., Bjerg, P.L., & Baun, A. 2010. Environmental benefits and risks of zero-valent iron nanoparticles (nZVI) for in situ remediation: Risk mitigation or trade-off? Journal of Contaminant Hydrology 118(3-4): 165-183.

Griffitt, R.J., Luo, J., Gao, J., Bonzongo, J., & Barber, D.S. 2008. Effects of particle composition and species on toxicity of metallic nanomaterials in aquatic organisms. Environmental Toxicology and Chemistry 27(9): 1972-1978.

Hall, S., Bradley, T., Moore, J.T., Kuykindall, T., & Minella, L. 2009. Acute and chronic toxicity of nano-scale TiO2 particles to freshwater fish, cladocerans, and green algae, and effects of organic and inorganic substrate on TiO2 toxicity. Nanotoxicology 3(2): 91-97.

Hang, J., Shi, L., Feng, X., & Xiao, L. 2009. Electrostatic and electrosteric stabilization of aqueous suspensions of barite nanoparticles. Powder Technology 192(2): 166-170.

Hansen, S.F., Michelson, E., Kamper, A., Borling, P., Stuer-Lauridsen, F. & Baun, A. 2008. Categorization framework to aid exposure assessment of nanomaterials in consumer products. Ecotoxicology 17(5): 438-447.

Hartmann, N.B., & Baun, A. 2010. The nano cocktail: Ecotoxicological effects of engineered nanoparticles in chemical mixtures. Integrated Environmental Assessment and Management 6(2): 311-313.

Hartmann, N.B., Buendia, I., Baun, A., 2011a. Biodegradation of aged aqueous suspensions of C60. Submitted to Environmental Pollution. Under revision.

Hartmann, N.B., Engelbrekt, C., Zhang, J., Ulstrup, J., Kusk, K.O., Baun, A., 2011b. The challenges of testing insoluble metal and metal oxide nanoparticles in algal bioassays: titanium dioxide and gold nanoparticles as case studies. In preparation.

Hartmann, N.B., Legros, S., von der Kammer, F., Hofmann, T., & Baun, A., 2011c. TiO2 nanoparticles as carriers for cadmium in Lumbriculus variegatus and Daphnia magna. In preparation.

Hartmann, N.B., von der Kammer, F., Hofmann, T., Baalousha, M., Ottofuelling, S. 2010. Algal testing of titanium dioxide nanoparticles - Testing considerations, inhibitory effects and modification of cadmium bioavailability. Toxicology 269(2-3): 190-197.

Hasler, A.D. 1935. The Physiology of Digestion of Plankton crustacea, I. Some digestive enzymes of daphnia. Biological Bulletin 68(2): 207-214.

61

Heinlaan, M., Ivask, A., Blinova, I., Dubourguier, H., & Kahru, A. 2008. Toxicity of nanosized and bulk ZnO, CuO and TiO2 to bacteria Vibrio fischeri and crustaceans Daphnia magna and Thamnocephalus platyurus. Chemosphere 71(7): 1308-1316.

Henry, T.B., Menn, F.M., Fleming, J.T., Wilgus, J., Compton, R.N., & Sayler, G.S. 2007. Attributing effects of aqueous C60 nano-aggregates to tetrahydrofuran decomposition products in larval zebrafish by assessment of gene expression. Environmental Health Perspectives 115(7): 1059-1065.

Hund-Rinke, K., & Simon, M. 2006. Ecotoxic effect of photocatalytic active nanoparticles (TiO2) on algae and daphnids. Environmental Science and Pollution Research - International 13(4): 225-232.

Hund-Rinke, K., Schlich, K., & Wenzel, A. 2010. TiO2 nanoparticles – Relationship between dispersion preparation method and ecotoxicity in the algal growth test. Umweltwissenschaften und Schadstoff-Forschung 22(5): 517-528.

Hutter, E., Boridy, S,, Labrecque, S., Lalancette-Hébert, M., Kriz, J., Winnik, F.M., & Maysinger, D. 2010. Microglial response to gold nanoparticles. ACS Nano 4(5): 2595-2606.

International Organisation for Standardisation (ISO). 2006. Water quality — Guidelines for algal growth inhibition tests with poorly soluble materials, volatile compounds, metals and waste water. ISO 14442:2006. International Organisation for Standardisation, Geneva, Switzerland

International Organisation for Standardisation (ISO). 2008. Nanotechnologies – Terminology and definitions for nano-objects – Nanoparticle, nanofibre and nanoplate. ISO/TS 27687. International Organisation for Standardisation, Geneva, Switzerland

Ivask, A., Bondarenko, O., Jepihhina, N., & Kahru, A. 2010. Profiling of the reactive oxygen species-related ecotoxicity of CuO, ZnO, TiO2, silver and fullerene nanoparticles using a set of recombinant luminescent Escherichia coli strains: differentiating the impact of particles and solubilised metals. Analytical and Bioanalytical Chemistry 398(2): 701-716.

Jafvert, C.T., & Kulkarni, P.P. 2008. Buckminsterfullerene's (C60) octanol-water partition coefficient (Kow) and aqueous solubility. Environmental Science & Technology 42(16): 5945-5950.

Johnston, B.D., Scown, T.M., Moger, J., Cumberland, S.A., Baalousha, M., Linge, K., van

Aerle, R., Jarvis, K., Lead, J.R., & Tyler, C.R. 2010. Bioavailability of nanoscale metal oxides TiO(2), CeO(2), and ZnO to fish. Environmental Science & Technology 44(3): 1144-1151.

Kahru, A., & Dubourguier, H. 2010. From ecotoxicology to nanoecotoxicology. Toxicology 269(2-3): 105-119.

Kahru, A., Dubourguier, H., Blinova, I., Ivask, A., & Kasemets, K. 2008. Biotests and biosensors for ecotoxicology of metal oxide nanoparticles: A minireview. Sensors 8(8): 5153-5170.

62

Kang, S., Goyal, A., Li, J., Gapud, A.A., Martin, P.M., Heatherly, L., Thompson, J.R., Christen, D.K., List, F.A., Paranthaman, M., & Lee, D.F. 2006. High-performance high-Tc superconducting wires. Science 311(5769): 1911 -1914.

Keller, A. A., Wang, H., Zhou, D., Lenihan, H.S., Cherr, G., Cardinale, B.J., Miller, R., & Ji, Z. 2010. Stability and aggregation of metal oxide nanoparticles in natural aqueous matrices. Environmental Science & Technology 44(6): 1962-1967.

Kennedy, A.J., Gunter, J.C., Chappell, M.A., Goss, J.D., Hull, M.S., Kirgan, R.A., & Steevens, J.A. 2009. Influence of nanotube preparation in aquatic bioassays. Environmental Toxicology and Chemistry 28(9): 1930-1938.

Kim, K., Edgington, A.J., Klaine, S.J., Cho, J., & Kim, S.D. 2009. Influence of multiwalled carbon nanotubes dispersed in natural organic matter on speciation and bioavailability of copper. Environmental Science & Technology 43(23): 8979-8984.

Kim, K.T., Klaine, S.J., Lin, S., Ke, P.C., & Kim, S.D. 2010. Acute toxicity of a mixture of copper and single-walled carbon nanotubes to Daphnia magna. Environmental Toxicology and Chemistry 29(1): 122-126.

Klaine, S.J., Alvarez, P.J., Batley, G.E., Fernandes, T.F., Handy, R.D., Lyon, D.Y., Mahendra, S., McLaughlin, M.J., & Lead, J.R. 2008. Nanomaterials in the environment: behavior, fate, bioavailability, and effects. Environmental Toxicology and Chemistry 27(9): 1825-1851.

Komatsu, K. 2005. First chemical synthesis of H2@C60 H. Kuzmany, J. Fink, M. Mehring, & S. Roth (eds). AIP Conference Proceedings 786(1): 50-55.

Kortenkamp, A, Backhaus, T , Faust M. 2009. State of the art report on mixture toxicity. Report to the European Commission (D-G Environment). Available at: http://ec.europa.eu/environment/chemicals/pdf/report_Mixture%20toxicity.pdf [Accessed February 10, 2011].

Koukal, B., Rossé, P., Reinhardt, A., Ferrari, B., Wilkinson, K.J., Loizeau, J.L., & Dominik, J. 2007. Effect of Pseudokirchneriella subcapitata (Chlorophyceae) exudates on metal toxicity and colloid aggregation. Water Research 41(1): 63-70.

Kreyling, W.G., Semmler-Behnke, M., & Chaudhry, Q. 2010. A complementary definition of nanomaterial. Nano Today 5(3): 165-168.

Kroto, H.W., Allaf, A.W., & Balm, S.P. 1991. C60: Buckminsterfullerene. Chemical Reviews 91(6): 1213-1235.

Lawrence, M.A.M., Davies, N.A., Edwards, P.A., Taylor, M.G., & Simkiss, K. 2000. Can adsorption isotherms predict sediment bioavailability? Chemosphere 41(7): 1091-1100.

Li, T., Albee, B., Alemayehu, M., Diaz, R., Ingham., L., Kamal., S., Rodriguez., & M., Bishnoi, S.W. 2010. Comparative toxicity study of Ag, Au, and Ag-Au bimetallic nanoparticles on Daphnia magna. Analytical and Bioanalytical Chemistry 398(2): 689-700.

Lin, D., Tian, X., Wu, F., & Xing, B. 2010. Fate and transport of engineered nanomaterials in the environment. Journal of Environmental Quality 39(6): 1896-1908.

http://ec.europa.eu/environment/chemicals/pdf/report_Mixture%20toxicity.pdf�

63

Liu, J., & Hurt, R.H. 2010. Ion release kinetics and particle persistence in aqueous nano-silver colloids. Environmental Science & Technology. Available at: http://www.ncbi.nlm.nih.gov/pubmed/20175529 [Accessed February 17, 2011].

Lovern, S.B., & Klaper, R. 2006. Daphnia magna mortality when exposed to titanium dioxide and fullerene (C60) nanoparticles. Environmental Toxicology and Chemistry 25(4): 1132-1137.

Lovern, S.B., Owen, H.A., & Klaper, R. 2008. Electron microscopy of gold nanoparticle intake in the gut of Daphnia magna. Nanotoxicology 2(1): 43-48.

Lovern, S.B., Strickler, J.R., & Klaper, R. 2007. Behavioral and physiological changes in Daphnia magna when exposed to nanoparticle suspensions (titanium dioxide, nano-C-60, and C(60)HxC(70)Hx). Environmental Science & Technology 41(12): 4465-4470.

Mackay, D., & Fraser, A. 2000. Bioaccumulation of persistent organic chemicals: mechanisms and models. Environmental Pollution 110(3): 375-391.

Maness, P.C., Smolinski, S, Blake, D.M., Huang, Z., Wolfrum, E.J., & Jacoby, W.A. 1999. Bactericidal activity of photocatalytic TiO2 reaction: toward an understanding of its killing mechanism. Applied and Environmental Microbiology 65(9): 4094-4098.

Menard, A., Drobne, D., & Jemec, A. 2011. Ecotoxicity of nanosized TiO2. Review of in vivo data. Environmental Pollution 159(3): 677-684.

Miao, A-J., Luo, Z., Chen, C-S., Chin, W.C., Santschi, P.H. & Quigg, A. 2010. Intracellular Uptake: A Possible Mechanism for Silver Engineered Nanoparticle Toxicity to a Freshwater Alga Ochromonas danica. PLoS ONE 5(12): e15196.

Miao, A-J., Schwehr, K.A., Xu, C. Zhang, S-J., Luo, Z., Quigg, A., & Santschi, P.H. 2009. The algal toxicity of silver engineered nanoparticles and detoxification by exopolymeric substances. Environmental Pollution 157(11): 3034-3041.

Miño, L.A., Folco, S., D’Ángelo, A.M.P.D., & Guerrero, N.R.V. 2006. Modeling lead bioavailability and bioaccumulation by Lumbriculus variegatus using artificial particles. Potential use in chemical remediation processes. Chemosphere 63(2): 261-268.

Miyauchi, M., Nakajima, A., Watanabe, T., & Hashimoto, K. 2002. Photocatalysis and Photoinduced Hydrophilicity of Various Metal Oxide Thin Films. Chemistry of Materials 14(6): 2812-2816.

Mouchet, F., Landois, P., Flahaut, E., Pinelli, E., & Gauthier, L. 2009. Carbon nanotubes in the environment and their potential ecotoxicity: Context and state of the art. Environnement, Risques & Santé. 8(1), 47-55.

Nam, J., Won, N., Jin, H., Chung, H., & Kim, S. 2009. pH-Induced aggregation of gold nanoparticles for photothermal cancer therapy. Journal of the American Chemical Society 131(38): 13639-13645.

Nanotechnology Industries Association (NIA). 2009. European Parliament approves labelling and notification of nanomaterials in cosmetics ). Available at: http://www.nanotechia.org/news/global/european-parliament-approves-labelling-and-notific [Accessed February 10, 2011].

http://www.ncbi.nlm.nih.gov/pubmed/20175529�

http://www.nanotechia.org/news/global/european-parliament-approves-labelling-and-notific�

http://www.nanotechia.org/news/global/european-parliament-approves-labelling-and-notific�

64

National Nanotechnology Initiative. 2007. Big things from a tiny world. Available at: http://www.nano.gov/Nanotechnology_BigThingsfromaTinyWorldspread.pdf [Accessed February 17, 2011].

Navarro, E., Baun, A., Behra, R., Hartmann, N.B., Filser, J., Miao, A.J., Quigg, A., Santschi, P.H., & Sigg, L. 2008. Environmental behavior and ecotoxicity of engineered nanoparticles to algae, plants, and fungi. Ecotoxicology 17(5): 372-386.

Navarro, E., Piccapietra, F., Wagner, B., Marconi, F., Kaegi, R., Odzak, N., Sigg, L., & Behra, R. 2008. Toxicity of silver nanoparticles to Chlamydomonas reinhardtii. Environmental Science & Technology 42(23): 8959-8964.

Nel, A.E., Mädler, L., Velegol, D., Xia, T., Hoek, E.M.V., Somasundaran, P., Klaessig, F., Castranova, V., & Thompson, M. 2009. Understanding biophysicochemical interactions at the nano-bio interface. Nat Mater 8(7): 543-557.

Neouze, M., & Schubert, U. 2008. Surface modification and functionalization of metal and metal oxide nanoparticles by organic ligands. Monatshefte für Chemie - Chemical Monthly 139(3): 183-195.

Newman, M.C., & Unger, M.A. 2002. Fundamentals of Ecotoxicology, 2nd ed. CRC Press, Boca Raton, Florida, United States. p.58-61.

Nowack, B., & Bucheli, T.D. 2007. Occurrence, behavior and effects of nanoparticles in the

environment. Environmental Pollution (Barking, Essex: 1987) 150(1): 5-22.

Oberdörster, E. 2004. Manufactured nanomaterials (fullerenes, C60) induce oxidative stress in the brain of juvenile largemouth bass. Environmental Health Perspectives 112(10): 1058-1062.

Oberdörster, E., Zhu, S., Blickley, T.M., McClellan-Green, P., & Haasch, M.L. 2006. Ecotoxicology of carbon-based engineered nanoparticles: Effects of fullerene (C60) on aquatic organisms. Carbon 44(6): 1112-1120.

Oberdörster, G., Oberdörster, E., & Oberdörster, J. 2005. Nanotoxicology: an emerging discipline evolving from studies of ultrafine particles. Environmental Health Perspectives 113(7): 823-839.

Oberdörster, G., Stone, V., & Donaldson, K. 2007. Toxicology of nanoparticles: A historical perspective. Nanotoxicology 1(1): 2-25.

Organisation for Economic Co-operation and Development (OECD). 2000. Guidance Document on Aquatic Toxicity Testing of Difficult Substances and Mixtures. OECD Series on Testing and Assessment No. 23. Available at: http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2000)6&doclanguage=en [Accessed February 10, 2011].

Organisation for Economic Co-operation and Development (OECD). 2001. Guidance Document on Transformation/Dissolution of Metals and Metal Compounds in Aqueous Media. Chemicals Testing Monographs No. 29. Available at: http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2001)9&doclanguage=en [Accessed February 10, 2011].

http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2000)6&doclanguage=en�




65

Organisation for Economic Co-operation and Development (OECD). 2004. Test No. 202: Daphnia sp. Acute Immobilisation Test

Organisation for Economic Co-operation and Development (OECD). 2008: Test No. 211: Daphnia magna Reproduction Test

Organisation for Economic Co-operation and Development (OECD). 2009. Preliminary Review of OECD Test Guidelines for their Applicability to Manufactured Nanomaterials. OECD Series on of Safety of Manufactured Nanomaterials No. 15. Available at: http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2009)21&doclanguage=en [Accessed February 10, 2011].

Organisation for Economic Co-operation and Development (OECD). 2010a. Guidance Manual for the Testing of Manufactured Nanomaterials: OECD Sponsorship Programme: First Revision. OECD Series on of Safety of Manufactured Nanomaterials No. 25. Available at: http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2009)20/rev&doclanguage=en [Accessed February 10, 2011].

Organisation for Economic Co-operation and Development (OECD). 2010b. Preliminary Guidance Notes on Sample Preparation and Dosimetry for the Safety Testing of Manufactured Nanomaterials. OECD Series on of Safety of Manufactured Nanomaterials No. 24. Available at: http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=ENV/JM/MONO(2010)25&doclanguage=en [Accessed February 10, 2011].

Organisation for Economic Co-operation and Development (OECD). 2010c. Sponsorship Programme for the Testing of Manufactured Nanomaterials. Available at: http://www.oecd.org/document/47/0,3746,en_2649_37015404_41197295_1_1_1_1,00.html [Accessed February 10, 2011].

Organisation for Economic Co-operation and Development (OECD). 2010d. List of Manufactured Nanomaterials and List of Endpoints for Phase One of the Sponsorship Programme for the Testing of Manufactured Nanomaterials: Revision. OECD Series on of Safety of Manufactured Nanomaterials No. 27. Available at: http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono%282010%2946&doclanguage=en [Accessed February 10, 2011].

Organisation for Economic Co-operation and Development (OECD). 2011. Chemicals Testing - Guidelines: About. Available at: http://www.oecd.org/about/0,3347,en_2649_34377_1_1_1_1_1,00.html [Accessed February 10, 2011].

Oughton, D.H., Hertel-Aas, T., Pellicer, E., Mendoza, E., & Joner, E.J. 2008. Neutron activation of engineered nanoparticles as a tool for tracing their environmental fate and uptake in organisms. Environmental Toxicology and Chemistry 27(9): 1883-1887.

Park, J., Henry, T.B., Menn, F., Compton, R.N., & Sayler, G. 2010. No bioavailability of 17α-ethinylestradiol when associated with nC60 aggregates during dietary exposure in adult male zebrafish (Danio rerio). Chemosphere 81(10): 1227-1232.

Penn, R.L., Tanaka, K., & Erbs, J. 2007. Size dependent kinetics of oriented aggregation. Journal of Crystal Growth 309(1): 97-102.



http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2009)20/rev&doclanguage=en�

http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2009)20/rev&doclanguage=en�

http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=ENV/JM/MONO(2010)25&doclanguage=en�

http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=ENV/JM/MONO(2010)25&doclanguage=en�

http://www.oecd.org/document/47/0,3746,en_2649_37015404_41197295_1_1_1_1,00.html�

http://www.oecd.org/document/47/0,3746,en_2649_37015404_41197295_1_1_1_1,00.html�

http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono%282010%2946&doclanguage=en�

http://www.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono%282010%2946&doclanguage=en�

http://www.oecd.org/about/0,3347,en_2649_34377_1_1_1_1_1,00.html�

66

Pérez, S., Farré, M., & Barceló, D. 2009. Analysis, behavior and ecotoxicity of carbon-based nanomaterials in the aquatic environment. TrAC Trends in Analytical Chemistry 28(6): 820-832.

Petersen, E.J., Akkanen, J., Kukkonen, J.V.K., & Weber, W.J. 2009. Biological uptake and depuration of carbon nanotubes by Daphnia magna. Environmental Science & Technology 43(8): 2969-2975.

Petersen, E.J., Huang, Q., & Weber, W.J. 2008. Ecological uptake and depuration of carbon nanotubes by Lumbriculus variegatus. Environmental Health Perspectives 116(4): 496-500.

Petersen, E.J., Huang, Q., & Weber, W.J. 2010. Relevance of octanol-water distribution measurements to the potential ecological uptake of multi-walled carbon nanotubes. Environmental Toxicology and Chemistry: n/a-n/a.

Pinna, N., & Niederberger, M. 2008. Oxide synthesis as cornerstone of nanoscience: Eur. J.

Inorg. Chem. 6/2008. European Journal of Inorganic Chemistry 2008(6): 825-825.

Poland, C.A., Duffin, R., Kinloch, I., Maynard, A., Wallace, W.A., Seaton, A., Stone, V., Brown, S., Macnee, W., & Donaldson, K. 2008. Carbon nanotubes introduced into the abdominal cavity of mice show asbestos-like pathogenicity in a pilot study. Nat Nano 3(7): 423-428.

Popov, V.N. 2004. Carbon nanotubes: properties and application. Materials Science and Engineering: R: Reports 43(3): 61-102.

Poynton, H.C., Lazorchak, J.M., Impellitteri, C.A., Smith, M.E., Rogers, K., Patra, M., Hammer, K.A., Allen, H.J., & Vulpe, C.D. 2011. Differential gene expression in Daphnia magna suggests distinct modes of action and bioavailability for ZnO nanoparticles and Zn ions. Environmental Science & Technology 45(2): 762-768.

Renault, S., Baudrimont, M., Mesmer-Dudons, N., Gonzalez, P., Mornet, S., & Brisson, A. et al. 2008. Impacts of gold nanoparticle exposure on two freshwater species: a phytoplanktonic alga (Scenedesmus subspicatus) and a benthic bivalve (Corbicula fluminea). Gold Bulletin 41(2): 116-126.

Roberts, A.P., Mount, A.S., Seda, B., Souther, J., Qiao, R., Lin, S., Ke, P.C., Rao, A.M., & Klaine, S.J. 2007. In vivo biomodification of lipid-coated carbon nanotubes by Daphnia magna. Environmental Science & Technology 41(8): 3025-3029.

Rodea-Palomares, I., Boltes, K., Fernández-Piñas, F., Leganés, F., García-Calvo, E., Santiago, J., & Rosal, R. 2011. Physicochemical characterization and ecotoxicological assessment of CeO2 nanoparticles using two aquatic microorganisms. Toxicological Sciences: An Official Journal of the Society of Toxicology 119(1): 135-145.

Rogers, N.J., Franklin, N.M., Apte, S.C., Batley, G.E., Angel, B.M., Lead, J.R, & Baalousha, M. 2010. Physico-chemical behaviour and algal toxicity of nanoparticulate CeO2 in freshwater. Environ. Chem. 7(1): 50-60.

Rosenkranz, P., Chaudhry, Q., Stone, V., & Fernandes, T.F. 2009. A comparison of nanoparticle and fine particle uptake by Daphnia magna. Environmental Toxicology and Chemistry 28(10): 2142-2149.

67

Salager, J. 1994. Interfacial Phenomena in Dispersed Systems. FIRP Booklet # E120-N, Version # 1 (1993) Translation (1994). Available at: http://nanoparticles.org/pdf/Salager-E120N.pdf [Accessed February 10, 2011].

Sayes, C.M., Wahi, R., Kurian, P.A., Liu, Y., West, J.L., Ausman, K.D., Warheit, D.B., & Colvin, V.L. 2006. Correlating nanoscale titania structure with toxicity: a cytotoxicity and inflammatory response study with human dermal fibroblasts and human lung epithelial cells. Toxicological Sciences: An Official Journal of the Society of Toxicology 92(1): 174-185.

Schreiner, K.M., Filley, T.R., Blanchette, R.A., Bowen, B.B., Bolskar, R.D., Hockaday, W.C., Masiello, C.A., & Raebiger, J.W. 2009. White-rot basidiomycete-mediated decomposition of C60 fullerol. Environmental Science & Technology 43(9): 3162-3168.

Schwarzenbach, R.P., Gschwend, P.M., & Imboden, D.M. 2003. Environmental organic chemistry. John Wiley and Sons, New York, United States (ISBN: 9780471357506). p.280

Scientific Committee on Emerging and Newly Identified Health Risks (SCENIHR). 2006. The appropriateness of existing methodologies to assess the potential risks associated with engineered and adventitious products of nanotechnologies Available at: http://ec.europa.eu/health/ph_risk/committees/04_scenihr/docs/scenihr_o_003b.pdf [Accessed February 10, 2011].

Ševčík, P., Čík, G., & Šeršen, F. 2009. Inhibition of toxic effects of chlorophenols on the growth of Chlorella vulgaris by modified TiO2 photocatalyst. Fresenius Environmental Bulletin 18 (11) 2165-2169

Sharma, V. 2009. Aggregation and toxicity of titanium dioxide nanoparticles in aquatic environment—A Review. Journal of Environmental Science and Health, Part A 44(14): 1485-1495.

Sijm, D.T.H.M., Rikken, M.G.J., Rorije, E., Traas, T.P., Mclachlan, M.S., & .Peijnenburg, W.J.G.M. 2007. Transport, accumulation and transformation processes. In: van Leeuwen, C., & Vermeire, T. eds. 2007. Risk Assessment of Chemicals. Dordrecht: Springer Netherlands. Pages 73-158

Spohn, P., Hirsch, C., Hasler, F., Bruinink, A., Krug, H.F., & Wick, P. 2009. C60 fullerene: A powerful antioxidant or a damaging agent? The importance of an in-depth material characterization prior to toxicity assays. Environmental Pollution 157(4): 1134-1139.

Stevens, M.M., & George, J.H. 2005. Exploring and engineering the cell surface interface. Science 310(5751): 1135-1138.

Stone, V., Hankin, S., Aitken, R., Aschberger, K., Baun, Anders, Christensen, F., Fernandes, T., Hansen, S.F., Hartmann, N.B., Hutchinson, G., Johnston, H., Micheletti, C., Peters, S., Ross, B., Sokull-Kluettgen, B., Stark, D., Tran, L., 2010. Engineered nanoparticles: Review of health and environmental safety (ENRHES). Project Final Report, European Commission, FP7 CSA #21843.

Stumm, W., & Morgan, J.J. 1996. Aquatic Chemistry 3rd ed., John Wiley and Sons, New York, United States (ISBN: 9780471511847). p.819.

http://nanoparticles.org/pdf/Salager-E120N.pdf�

http://ec.europa.eu/health/ph_risk/committees/04_scenihr/docs/scenihr_o_003b.pdf�

68

Sun, H., Zhang, X., Niu, Q., Chen, Y., & Crittenden, J.C. 2006. Enhanced accumulation of arsenate in carp in the presence of titanium dioxide nanoparticles. Water, Air, and Soil Pollution 178(1-4): 245-254.

Sun, H., Zhang, X., Zhang, Z., Chen, Y., & Crittenden, J.C. 2009. Influence of titanium dioxide nanoparticles on speciation and bioavailability of arsenite. Environmental Pollution 157(4): 1165-1170.

Syberg, K., Jensen, T.S., Cedergreen, N., & Rank, J. 2009. On the use of mixture toxicity assessment in REACH - and the water framework directive: A review. Human and Ecological Risk Assessment: An International Journal 15(6): 1257.

Templeton, R.C., Ferguson, P.L., Washburn, K.M., Scrivens, W.A., & Chandler, G.T. 2006. Life-cycle effects of single-walled carbon nanotubes (SWNTs) on an estuarine meiobenthic copepod. Environmental Science & Technology 40(23): 7387-7393.

Tervonen, K., Waissi, G., Petersen, E.J., Akkanen, J., & Kukkonen, J.V. 2010. Analysis of fullerene-C60 and kinetic measurements for its accumulation and depuration in Daphnia magna. Environmental Toxicology and Chemistry 29(5) 1072-1078.

Thakkar, K.N., Mhatre, S.S., & Parikh, R.Y. 2010. Biological synthesis of metallic nanoparticles. Nanomedicine: Nanotechnology, Biology and Medicine 6(2): 257-262.

The Project on Emerging Nanotechnologies. Analysis • Consumer Products • Nanotechnology Project. Available at: http://www.nanotechproject.org/inventories/consumer/analysis_draft/ [Accessed February 10, 2011].

Tiede, K., Hassellöv, M., Breitbarth, E., Chaudhry, Q., & Boxall, A.B.A. 2009. Considerations for environmental fate and ecotoxicity testing to support environmental risk assessments for engineered nanoparticles. Journal of Chromatography. A 1216(3): 503-509.

United States Environmental Protection Agency (US EPA). 2003. Title 40 of the Code of Federal Regulations, Chapter I, Subchapter D, Part 136: Guidelines Establishing Test Procedures for the Analysis of Pollutants; Whole Effluent Toxicity Test Methods, Appendix C, page 320 Available at: http://www.epa.gov/earth1r6/6pd/qa/qadevtools/mod4references/analytical_references/40cfr136_03.pdf [Accessed February 10, 2011].

Van Geest, J.L., Poirier, D.G., Sibley, P.K., & Solomon, K.R. 2010. Measuring bioaccumulation of contaminants from field-collected sediment in freshwater organisms: A critical review of laboratory methods. Environmental Toxicology and Chemistry 29(11): 2391-2401.

Van Hoecke, K., De Schamphelaere, K.A.C., Van der Meeren, P., Lucas, S., & Janssen, C.R. 2008. Ecotoxicity of silica nanoparticles to the green alga Pseudokirchneriella subcapitata: importance of surface area. Environmental Toxicology and Chemistry / 27(9): 1948-1957.

http://www.nanotechproject.org/inventories/consumer/analysis_draft/�

http://www.epa.gov/earth1r6/6pd/qa/qadevtools/mod4references/analytical_references/40cfr136_03.pdf�

http://www.epa.gov/earth1r6/6pd/qa/qadevtools/mod4references/analytical_references/40cfr136_03.pdf�

69

Van Hoecke, K., Quik, J.T., Mankiewicz-Boczek, J., De Schamphelaere, K.A., Elsaesser, A., Van der Meeren, P., Barnes, C., McKerr, G., Howard, C.V., Van de Meent, D., Rydzyński, K., Dawson, K.A., Salvati, A., Lesniak, A., Lynch, I., Silversmit, G., De Samber, B., Vincze, L. & Janssen, C.R. 2009. Fate and effects of CeO2 nanoparticles in aquatic ecotoxicity tests. Environmental Science & Technology 43(12): 4537-4546.

von der Kammer, F., Ottofuelling, S., & Hofmann, T. 2010. Assessment of the physico-chemical behavior of titanium dioxide nanoparticles in aquatic environments using multi-dimensional parameter testing. Environmental Pollution 158(12): 3472-3481.

Warheit, D.B., Borm, P.J.A., Hennes, C., & Lademann, J. 2007. Testing strategies to establish the safety of nanomaterials: conclusions of an ECETOC workshop. Inhalation Toxicology 19(8): 631-643.

Wei, C., Zhang, Y., Guo, J., Han, B., Yang, X., & Yuan, J. 2010. Effects of silica nanoparticles on growth and photosynthetic pigment contents of Scenedesmus obliquus. Journal of Environmental Sciences 22(1): 155-160.

Worms, I., Simon, D.F., Hassler, C.S., & Wilkinson, K.J. 2006. Bioavailability of trace metals to aquatic microorganisms: importance of chemical, biological and physical processes on biouptake. Biochimie 88(11): 1721-1731.

Xia, X., Monteiro-Riviere, N.A., & Riviere, J.E. 2010. An index for characterization of nanomaterials in biological systems. Nature Nanotechnology 5(9): 671-675.

Yabe, S., & Sato, T. Cerium oxide for sunscreen cosmetics. Journal of Solid State Chemistry 171(1-2): 7-11.

Yeung, K.L., Leung, W.K., Yao, N., & Cao, S. 2009. Reactivity and antimicrobial properties of nanostructured titanium dioxide. Catalysis Today 143(3-4): 218-224.

Zhang, H., Penn, R.L., Hamers, R.J., & Banfield, J.F. 1999. Enhanced adsorption of molecules on surfaces of nanocrystalline particles. The Journal of Physical Chemistry B 103(22): 4656-4662.

Zhang, X., Sun, H., Zhang, Z., Niu, Q., Chen, Y., & Crittenden, J.C. 2007. Enhanced bioaccumulation of cadmium in carp in the presence of titanium dioxide nanoparticles. Chemosphere 67(1): 160-166.

Zhao, C., & Wang, W. 2010. Biokinetic uptake and efflux of silver nanoparticles in Daphnia magna. Environmental Science & Technology 44(19): 7699-7704.

Zhu, S., Oberdörster, E., & Haasch, M.L. 2006. Toxicity of an engineered nanoparticle (fullerene, C60) in two aquatic species, Daphnia and fathead minnow. Marine Environmental Research 62(Supplement 1): S5-S9.

Zhu, X., Chang, Y., & Chen, Y. 2010. Toxicity and bioaccumulation of TiO2 nanoparticle aggregates in Daphnia magna. Chemosphere 78(3): 209-215.

Zhu, X., Zhu, L., Chen, Y., & Tian, S. 2008. Acute toxicities of six manufactured nanomaterial suspensions to Daphnia magna. Journal of Nanoparticle Research 11(1): 67-75.

Zumdahl, S.S. 1998. Chemical Principles 3rd ed., Boston, United States (ISBN: 9780395839959). p.791-828

70

Žutić, V. & Svetličić, V. 2000. Interfacial processes. The Handbook of Environmental Chemistry, Vol. 5 Part D, Marine Chemistry, Springer-Verlag, Berlin Heidelberg, 149-165. Availavle at: http://www.irb.hr/-Adwp-/en/str/zimo/laboratoriji/lem/radovi./chapter.pdf/ [Accessed February 10, 2011].

71

Appendices

I. Hartmann, N.B., von der Kammer, F., Hofmann, T., Baalousha, M., Ottofuelling, S., & Baun, A., 2010. Algal testing of titanium dioxide nanoparticles – Testing considerations, inhibitory effects and modification of cadmium bioavailability. Toxicology, 269 (2-3) 190–197.

II. Hartmann, N.B. & Baun, A., 2010. The nano cocktail: ecotoxiological effects of engineered nanoparticles in chemical mixtures. Integrated Environmental Assessment and Management, 6 (2) 311–313.

III. Hartmann, N.B., Buendia, I., & Baun, A., 2011. Degradability of aged aqueous suspensions of C60 nanoparticles. Submitted to Environmental Pollution. Under revision.

IV. Hartmann, N.B., Engelbrekt, C., Zhang, J., Ulstrup, J., Kusk, K.O., & Baun, A., 2011. The challenges of testing insoluble metal and metal oxide nanoparticles in algal bioassays: titanium dioxide and gold nanoparticles as case studies, Manuscript

V. Hartmann, N.B., Legros, S., von der Kammer, F., Hofmann, T., & Baun, A., 2011. The potential of TiO2 nanoparticles as carriers for heavy metal uptake in Lumbriculus variegatus and Daphnia magna. Manuscript

VI. Baun, A., Hartmann, N.B., Grieger, K.D., & Kusk, K.O., 2008. Ecotoxicity of engineered nanoparticles to aquatic invertebrates: a brief review and recommendations for future toxicity testing. Ecotoxicology, 17 (5) 387-395.

VII. Baun, A., Sørensen, S.N., Rasmussen, R.F., Hartmann, N.B., & Koch, C.B., 2008, Toxicity and bioaccumulation of xenobiotic organic compounds in the presence of aqueous suspensions of aggregates of nano-C60. Aquatic Toxicology 86 (3) 379–387.

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