Surface reactivity, stability, and mobility of metal

nanoparticles in aqueous solutions – Influence of natural organic matter and implications

on particle dispersion preparation

Sulena Pradhan

Licentiate thesis

KTH Royal Institute of Technology

School of Chemical Science and Engineering Department of Chemistry

Division of Surface and Corrosion Science SE-100 44 Stockholm, Sweden

Stockholm, 2017

This licentiate thesis will, with the permission of Kungliga Tekniska Högskolan, Stockholm, be

presented at a licentiate seminar on September 21, 2017, at 10 a.m. in E 31 Lindstedtsvägen 3,

KTH Campus, Stockholm, Sweden.

Surface reactivity, stability, and mobility of metal

nanoparticles in aqueous solutions

– Influence of natural organic matter and implications

on particle dispersion preparation

TRITA-CHE Report 2017:33 ISSN 1654-1081 ISBN 1654 - 1081

KTH Royal Institute of Technology School of Chemical Science and Engineering Department of Chemistry

Division of Surface and Corrosion Science Drottning Kristinas väg 51

SE-100 44 Stockholm

Sweden

The following item is printed with permission:

Paper I: Journal of Nanoparticle Research, open access Denna uppsats är skyddad enligt upphovsrättslagen. Alla rättigheter förbehålls. Copyright © Sulena Pradhan 2017. All rights reserved. No part of this thesis may be reproduced

by any means without permission from the author.

LIST OF SCIENTIFIC PUBLICATIONS

Publication included in this thesis Paper I. Pradhan, Sulena; Hedberg, Jonas; Blomberg, Eva; Wold, Susanna; Odnevall

Wallinder, Inger. Effect of sonication on particle dispersion, administered dose and metal

release of non-functionalized, non-inert metal nanoparticles, Journal of Nanoparticle Research,

2016, 18, 285 Paper II. Pradhan, Sulena; Hedberg, Jonas; Rosenqvist, Jörgen; Jonsson, Caroline M. ; Wold,

Susanna; Blomberg Eva; Odnevall Wallinder, Inger. Influence of humic acid and dihydroxy

benzoic acid on the agglomeration, sedimentation and dissolution of copper, manganese,

aluminum and silica nanoparticles. Manuscript

Author’s contribution Paper I. Experimental work, data analysis and writing the first draft of the publication Paper II.

Major experimental work except for ATR–FTIR, data analysis except for ATR–FTIR and silica NP characterization and writing the first draft of the

publication

Other publications not within the framework of this thesis Hedberg, Yolanda; Pradhan, Sulena; Cappellini, Francesca; Karlsson, Maria-Elisa; Blomberg,

Eva; Karlsson, Hanna; Odnevall Wallinder, Inger; Hedberg, Jonas. Electrochemical surface

oxide characteristics of metal nanoparticles (Mn, Cu and Al) and the relation to toxicity,

Electrochimica Acta, 2016, 212, 360-371 Hedberg, Yolanda ; Pettersson, Maria ; Pradhan, Sulena ; Odnevall Wallinder, Inger ; Rutland,

Mark ; Persson, Cecilia. Can cobalt (II) and chromium (III) ions released from joint prostheses

influence the friction coefficient? ACS Biomaterials Science & Engineering, 2016, 1(8), 617-

620. Pradhan, Sulena; Hedberg, Jonas; Wold, Susanna; Blomberg, Eva; Odnevall Wallinder, Inger.

Implementation of the standard operation procedure used and stipulated in the NanoReg EU

project on preparation of nanoparticle dispersion for further analysis., Technical Report, Mistra

Environmental Nanosafety, Sept. 2015 Ferraz, Natalia; Strömme, Maria; Fellström, Bengt; Pradhan, Sulena; Nyholm Leif; Mihranyan,

Albert. In vitro and in vivo toxicity of rinsed and aged nanocellulose-polypyrrole Composites,

Journal of Biomedical Materials Research Part A, 2012, 100, 2128-2138

Johansson, Erik; Pradhan, Sulena; Wang, Engang; Unger, Eva; Hagfeldt, Anders; Mats R,

Efficient infiltration of low molecular weight polymer in nanoporous TiO2. Chemical Physics

Letters, 2010, 502, 225-230

CONFERENCES AND WORKSHOPS

1. Poster - Can organic degradation products stabilize metal nanoparticles in solution?

Sulena Pradhan, Jonas Hedberg, Eva Blomberg, Inger Odnevall Wallinder, Susanna

Wold 2

nd Nanosafety forum for young scientist, September 15-16, 2016, Visby, Sweden.

2. Mistra meetings – November 11-12, 2015, Gothenburg, Sweden.

November 24-25, 2016, Stockholm, Sweden.

3. KTH PhD day - Pitch presentation titled “When small meets smaller”

September 5th

, 2016, Lidingö, Sweden.

ABSTRACT

The growing development of nanotechnology has resulted in an increased use of nanoparticles

(NPs) in various applications ranging from medicine, military, to daily consumer products.

There is a concern that NPs can be dispersed into the environment in various ways, for example

to air and water during manufacture, use, incineration or recycling of products and thus pose a

risk to health and the environment. Risk assessments of NPs are hence necessary. One property

of NPs, which may make them very useful and at the same time potentially harmful, is their

small size (in nanometer range) and hence high surface area per NP mass. This study forms part of the National Mistra Environmental Nanosafety Research Program. The

program provides an interdisciplinary platform for researchers from e.g. nanoscience, medicine,

chemistry, material science, life cycle analysis, and social science. Specific aspects of this

program involve characterization of NPs in different environmental settings, toxicity studies of

aquatic organisms, integrated risk assessment of NPs, and societal dimensions of nanosafety.

The contribution of this thesis within the program includes studies of stability and mobility of

metal NPs and their extent of transformation/dissolution upon environmental interaction.

Environmental risk assessments of NPs require a detailed understanding of how they change in

terms of physical and chemical properties (charge, size, and surface oxide composition),

important aspects for their stability, mobility, and reactivity in the environment. Generated data

is highly relevant for the other activities of the Mistra Environmental Nanosafety program, e.g.

to gain an improved understanding and design of particle dispersions and ecotoxicity studies, as

any environmental interaction will result in the transformation/dissolution of the NPs and

change the surface chemistry (e.g. adsorption of natural organic matter, changes in surface oxide

properties), aspects that largely influence their speciation and potential toxicity.

Common sonication protocols exist to prepare particle dispersions for different in vitro studies.

The influence of key parameters stipulated by these protocols on the particle size,

transformation/dissolution, and extent of sedimentation was investigated for bare metal NPs.

Improved knowledge on these aspects is crucial for design and interpretation of results of NP-

related investigations. Reactive metal NPs such as Cu and Mn NPs started to dissolve and

release metals already during the probe sonication step of the stock solution, and that the

presence of bovine serum albumin (often added as a stabilizing agent) enhanced this process.

Even though prolonged sonication time i.e. increased delivered acoustic energy, reduced the size

of formed agglomerates, sedimentation was still significant. As a consequence, administered

doses from pipetted stock solutions were significantly lower (30-70%) than the nominal doses.

The main reason behind the significant extent of agglomeration, with concomitant

sedimentation, is related to the strong van der Waals forces prevailing between metal NPs. It is

hence essential to determine the administrated dose of metallic NPs in e.g. nanotoxicological

testing. Interactions between metallic NPs and natural organic matter (NOM) were studied in terms of

stability, mobility and metal dissolution in order to mimic a potential exposure scenario. NOM

was represented by humic acid (HA), a main component of organic matter in the environment,

and by dihydroxybenzoic acid (DHBA), a small degradation product of NOM. Sedimentation of

the Cu, and the Al NPs were slower in the presence of NOM in freshwater compared with

freshwater only, whereas the effect of NOM was small for the Mn NPs. Stabilization was related

to surface adsorption of NOM, which increased the steric repulsion between the particles, and in

the case of HA also increased the magnitude of the zeta potential

(resulting in increased electrostatic repulsion). Slight initial increase in particle stability was

observed in freshwater containing DHBA, but after 24 h, sedimentation of the NPs was

comparable to the conditions in freshwater only. The presence of HA (at a concentration of 20

mg/L) was found to stabilize the NPs in freshwater for more than 24 h. However, both the lower

and higher HA concentration (2 and 40 mg/L) resulted in agglomeration of the Cu and Al NPs

already within a few hours. Mn NPs were more stable in terms of sedimentation in freshwater at

all three humic acid concentrations. This concludes that the concentration and type of NOM

largely influence the stability of the studied metal NPs in solution. In contrast, SiO2 NPs were

not influenced by the presence of NOM in terms of stability, most probably predominantly

related to smaller attractive van der Waals forces and larger electrostatic repulsion (due to

higher surface charge) compared with the metal NPs. Metal release from the Cu and Al NPs was enhanced in the presence of NOM, whereas no

significant influence was observed for the Mn NPs. All metal NPs were dissolved relatively

fast; 10% or more of the particle mass was dissolved within 24 h. Speciation predictions

revealed rapid complexation between released Cu and Al in solution and NOM, reducing the

bioavailability, whereas less complexation was evident for released Mn (as ions). In all, rapid

agglomeration and sedimentation imply that any risks associated with the environmental

dispersion of these metal NPs will be limited to the vicinity of their source. Mn NPs, having

lower sedimentation rates than the Cu and Al NPs, and lack of solution complexation of

released ions will likely have a relatively higher probability to be mobile and transported to

other aquatic settings.

SAMMANFATTNING

Nanoteknologins utveckling har resulterat i en ökad användning av nanopartiklar (NP) i olika

applikationer, t.ex. i läkemedel, militär utrustning och dagliga konsumentprodukter. Det finns

dock en oro för att NP kan spridas till miljön på olika sätt, t.ex. till luft och vatten vid

tillverkning, användning, förbränning eller återvinning av produkter och därmed utgöra en

miljö- och hälsorisk. Riskbedömningar av NP är därför nödvändiga att genomföra. En egenskap

hos NP, som både gör dem mycket användbara, och samtidigt potentiellt skadliga, är deras

storlek (i nanometer-nivå) och därmed stora ytarea. Denna studie utgör en del av det nationella forskningsprogrammet Mistra Environmental

Nanosafety, en tvärvetenskaplig plattform för forskare inom nanovetenskap, miljövetenskap,

medicin, ykemi, materialvetenskap, livscykelanalys och samhällsvetenskap. Specifika aspekter

som studeras i detta program omfattar bl.a. karakterisering av NP och hur deras egenskaper

förändras i olika vattenmiljöer, toxicitetsstudier av vattenlevande organismer, integrerad

riskbedömning av NPs och samhälleliga dimensioner av nanosäkerhet. Denna avhandling bidrar

till forskningen genom ingående studier av stabiliteten, mobiliteten och upplösningshastigheten

av metalliska NP i kontakt med olika vattensystem. Miljöriskbedömningar av NP kräver en

detaljerad förståelse av förändring av deras fysiskaliska och kemiska egenskaper (t.ex. laddning,

storlek, ytoxid-sammansättning), vilka påverkar deras stabilitet, rörlighet och reaktivitet i

miljön. Erhållna forskningsdata används även inom ramen för övriga verksamheter i

forskningsprogrammet, t.ex. för en bättre förståelse och design av de ekotoxikologiska

undersökningarna, eftersom växelverkan med miljön kan resultera i en total eller delvis

upplösning av NPs samt att förändringar i NPs ytkemi (t.ex. genom adsorption av naturligt

organiskt material, förändringar i ytoxidens egenskaper) i högsta grad påverkar deras reaktivitet

och potentiellt toxiska egenskaper. Inom nanotoxikologin finns olika protokoll att följa för att förbereda stamlösningar av NP

genom sonikering. Kunskap saknas dock till stor del om hur dessa riktlinjer påverkar storlek,

omvandling/upplösning, och grad av agglomeration och sedimentation hos reaktiva metalliska

NP. Dessa aspekter har undersökts i denna avhandling då de är av avgörande betydelse för

experimentell design och tolkning av resultat av NP-relaterade undersökningar. Resultaten visar

på snabb partikelagglomeration, sedimentation och frisättning av metaller från de reaktiva

partiklarna redan under prepareringen av stamlösningarna genom sonikering

(ultraljudsbehandling). Andelen frisatta metaller ökade dessutom i omfattning genom tillsats av

bovine serum albumin, ett protein som ofta tillsätts för att stabilisera partiklar i lösning. Även

om en förlängd ultraljudsbehandling, dvs. en större tillförd mängd akustisk energi, minskade

storleken på bildade agglomerat var sedimentationen fortfarande betydande. Dessa snabba

processer medför att de partikeldoser som från stamlösningen sedan tillförs t.ex. till celler i

nanotoxikologiska studier kan vara betydligt lägre (30-70% för Cu, Al, Mn NP i denna studie)

än de nominella doserna. Den främsta orsaken bakom den betydande omfattningen av

agglomeration och efterföljande sedimentation är relaterad till starka van der Waals-krafter som

råder mellan metalliska NPs. Resultaten visar tydligt att mätningar av tillförda doser måste

kvantifieras för olika system. Växelverkan mellan metalliska NP och naturligt organiskt material (NOM) studerades med

avseende på stabilitet, rörlighet och metallfrisättning för att efterlikna ett potentiellt

exponeringsscenario där partiklarna kommer i kontakt med t.ex. sötvatten. NOM representerades

i denna studie av humussyra (HA), en huvudkomponent i organiskt material i

miljön, samt av dihydroxybenzosyra (DHBA), som utgör små nedbrytningsprodukter av NOM.

Sedimentationen av agglomerat av Cu och Al NP var långsammare i närvaro av NOM jämfört

med syntetiskt ytvatten utan NOM. Den ökade partikelstabiliteten var för dessa partiklar

relaterad till ytadsorptionen av NOM, vilket ökade de sterisk krafterna mellan partiklarna och

också partiklarnas ytladdning. Motsvarande effekt för Mn NPs var dock liten trots NOM-

adsorption. En liten initial ökning av partiklarnas stabilitet observerades i ytvatten som innehöll

DHBA, men efter 24 h var sedimentation jämförbar med förhållandena i ytvatten utan NOM.

HA i en koncentration av 20 mg/L, realistisk för verkliga förhållanden, stabiliserade partiklarna

längre än 24 h. En både lägre och högre HA koncentration (2 och 40 mg/L) resulterade i

agglomeration av både Cu NP och Al NP redan inom några timmar medan Mn NP var mer

stabila för alla koncentrationer. Resultaten visar att både koncentrationen och typen av NOM

påverkar stabiliteten och mobiliteten av de metalliska NP i lösning. NP av SiO2 påverkades

däremot inte alls av närvaron av NOM med avseende på stabilitet, troligen huvudsakligen

beroende på lägre attraktiva van der Waals krafter och större elektrostatisk repulsion (på grund

av högre ytladdning) jämfört med de metalliska partiklarna. Andelen frisatta metaller från Cu och Al NP ökade i närvaro av NOM, medan ingen signifikant

effekt observerades för Mn NP. Samtliga NP upplöstes relativt snabbt, 10% eller mer av

partiklarnas massa inom 24 h. Beräkningar av specieringen av mängden frisatta metaller i

lösning visade på snabb komplexering mellan frisatt Cu och Al i ytvatten och NOM, vilket

pekar på låg biotillgänglighet, medan en lägre grad av komplexering med NOM predikterades

för frisläppt Mn. På grund av snabb agglomeration och sedimentation av de metalliska NPs

undersökta i denna studie förväntas eventuella risker i samband en spridningen av dessa

metalliska NPs begränsas till källans närhet. Mn NP, med lägre sedimentationshastighet än Cu

och Al NP, samt brist på komplexering av frisatta joner, har troligen en relativt högre

sannolikhet att vara mobila och kunna transporteras i miljön.

CONTENTS

1. AIM OF THIS WORK ........................................................................................................ 1

1.1 Motivation ..................................................................................................................... 1

1.2 Aim ............................................................................................................................... 2 2. INTRODUCTION ............................................................................................................... 2

2.1 Nanoparticles and nanomaterials..................................................................................... 2

2.2 Environmental concern ................................................................................................... 2

2.3 Risk assessment of NPs .................................................................................................. 3 3. BACKGROUND ................................................................................................................. 4

3.1 Choice of NPs ................................................................................................................ 4

3.2 Relevant NPs characteristics from an environmental perceptive ....................................... 5

3.3 Selected phenomena influencing NP fate in the environment ........................................... 6

3.3.1 Colloidal stability .................................................................................................... 6

3.3.2 Metal release/ dissolution ......................................................................................... 7 4. EXPERIMENTAL APPROACH ......................................................................................... 8

4.1 Sample preparation ........................................................................................................ 8

4.2 Characterizations of NPs ................................................................................................ 9

4.2.1 Size distribution ....................................................................................................... 9

4.2.2 Zeta potential of particles in dispersion ..................................................................... 9

4.2.3 Dissolution/metal release ....................................................................................... 11

4.2.4 Adsorption studies of organic matter ...................................................................... 11 5. KEY FINDINGS (Papers I and II)...................................................................................... 12

5.1 Sonication of NP suspensions largely influence the particle size and extent of metal release (Paper I) ................................................................................................................. 12

5.2 Rapid sedimentation of NPs in the stock solution results in significantly lower administered doses than the nominal doses of sonicated metal NP dispersions – an effect of strong attractive van der Waals forces (Paper I) .................................................................. 15

5.3 Metals are released from Cu, Al and Mn NPs already during the sonication step and the presence of biomolecules can enhance this process (Paper I) ............................................... 16

5.4 The extent of particle agglomeration of the Cu, Al and Mn NPs depends on the type of surface adsorbed NOM (Paper II) ....................................................................................... 17

5.5 NOM delayed the sedimentation of the Cu and Al NPs, in particular pronounced in the

presence of HA, whereas no significant effect was observed for the Mn NPs (Paper II). ... 19

5.6 The release of metals from the Cu, and Al NPs was enhanced in the presence of NOM, whereas no significant effect was observed for the Mn NPs despite NOM adsorption. Released Cu and Al readily formed strong NOM-complexes of low bioavailability in solution whereas released Mn remained as free ions (Paper II). ........................................... 20

6. MAIN MESSAGES .......................................................................................................... 22 7. FUTURE WORK ............................................................................................................. 23 8. ACKNOWLEDGMENTS ................................................................................................. 24 9. REFERENCES ................................................................................................................. 25

LIST OF ABBREVIATIONS

AAS ATR-FTIR

DHBA DLS DLVO HA NOM PCCS XPS

Atomic absorption spectroscopy Attenuated total reflectance Fourier transform infrared spectroscopy Dihydroxybenzoic acid Dynamic Light Scattering Derjaguin, Landau, Verwey and Overbeek Humic acid Natural organic matter Photon cross correlation spectroscopy X-ray photoelectron spectroscopy

1. AIM OF THIS WORK 1.1 Motivation The growing development and innovation of nanoparticles (NPs) for different applications

rapidly results in an increasing number of NP-based products [1] with metal NPs being an

important group of NPs [2, 3]. There is a concern that NPs (including metal NPs) may be

released into the environment during production, use, or disposal of these products. Other

important sources for metallic NPs include traffic emissions and wear particles from e.g. tires

and brake pads, reported as major respiratory pollutants in urban areas [4]. Due to their small

size, it has become important to assess possible risks of their environmental dispersion, as they

in some cases may be harmful to humans or aquatic organisms. Recent modern analytical

techniques have enabled investigations of different properties of NPs in the environment.

Nevertheless, there are still knowledge gaps concerning changes in physico-chemical

characteristics and transformation of metal NPs upon environmental entry [5]. The concern for adverse effects induced by NPs has resulted in the development of different

national and international research projects aiming for in-depth studies of different aspects

including e.g. NP toxicity, transformation/dissolution, environmental fate, and hazard

assessment. This thesis forms part of the national Mistra Environmental Nanosafety Program

initiated by the Swedish Foundation for Strategic Environmental Research (MISTRA). The

framework consists of different academic groups from Chalmers University of Technology,

Karolinska Institutet, KTH Royal Institute of Technology, Gothenburg University, and Lund

University, together with industrial partner Akzo Nobel. The program provides an

interdisciplinary platform for researchers from e.g. nanoscience, medicine, surface chemistry,

material science, life cycle analysis, and social science. The program focuses on

characterization of NPs in different environmental aquatic media, toxicity studies in the food

chain using different aquatic organisms, integrated risk assessment of NPs, and societal

dimensions of nanosafety. More specifically, the program is divided into five closely integrated

work packages (WP); 1) exposure, fate and life cycle assessments; 2) the molecular nano-

interface; 3) integrated hazard assessments; 4) social dimensions of nanosafety, and 5)

nanotechnological solutions to environmental problems. Three case studies are overarching

these work packages; i) emissions of NPs from automobile applications to road run-off water, ii) systematic studies of commercially available nanomaterials, and iii) future nanomaterials.

The Mistra Environmental Nanosafety Program supports and contributes to the Swedish

environmental quality objective “A non-toxic environment” with the objective to hand over an

environment with reduced chemical risks to the next generation. This thesis has been performed within the framework of WP 1 in the Mistra Environmental

Nanosafety program, with implications for case study 1, in which transformations of metal NPs

have been investigated. Data and knowledge are generated related to metal NP agglomeration,

sedimentation, transformation/dissolution, and changes in surface chemistry and speciation of

metal NPs in environmental-relevant media such as freshwater. These aspects are highly

important and integrated within the framework of the other work packages of the program, e.g.

used to understand and design relevant experiments on ecotoxicity, metal dissolution, and

changes in surface chemistry (adsorption of natural organic matter, surface oxide properties).

Another example is the generation of rates of agglomeration, sedimentation, and dissolution,

essential data in modelling of the fate and risks of NPs in the environment [6, 7]. In addition,

thorough work has been done to assess the influence of probe sonication, the

1

most common method used for preparation of NP dispersions, on the particle and surface

characteristics, dissolution properties and agglomerate size. These aspects are crucial as they

largely influence e.g. nanotoxicological interpretations. This work also relates to the 2030

agenda made by UN on global sustainability goals, Sep.2015 set to protect the planet from

different man-made obstacles for a peaceful living. Among the 17 sustainable goals, life below

water, life on land, and clean water and sanitation focus on minimizing the use of hazardous

chemicals. Since NPs and nanomaterials are the new age chemical substances, knowledge about

their potential risks on the environment is a very important for a safe planet and to achieve the

set goals. 1.2 Aim The aim of this work was to study metal NPs including mainly copper (Cu), aluminum (Al) and manganese (Mn) related to the following specific objectives:

1. To test and assess how the key parameters of an existing dispersion protocol for

sonication of NPs influence the stability/mobility, dissolution/transformation and

particle characteristics of reactive metal NPs, and propose relevant sonication settings

for reactive metal NPs. 2. To obtain an improved understanding of agglomeration, sedimentation and dissolution

processes of metal NPs in freshwater, as well as providing dissolution and sedimentation

rates. 3. To investigate how different types of natural organic matter (NOM) interact with metal

NPs from an adsorption, agglomeration, sedimentation and transformation/dissolution

perspective, and if possible deduce possible NP-specific interactions with NOM.

2. INTRODUCTION 2.1 Nanoparticles and nanomaterials The history of nanotechnology dates back to 600 BC when carbon nanotubes were used in

Wootz steel [8]. Earlier use of nanotechnology was also related to painting or art, until the

discovery of the modern electron microscope in 1931 [9]. Nanotechnology then became a

sophisticated science for materials in the scale of 1-100 nm. For NPs, the surface properties are

in general more important for the physical and chemical properties compared to the

corresponding bulk material characteristics [10]. For example, NPs used in photovoltaic cells

have higher absorbance capacity than if used within continuous sheets of bulk materials [11].

NPs and nanomaterials (NMs) possess large surface areas due to a large number of surface

atoms, a property that is useful in different industrial applications. Today, NPs are used in

various societal sectors such as medicine, military, astronomy, and also in consumer products

such as self-cleaning glass, cosmetics, detergents and antibacterial clothing [12]. 2.2 Environmental concern The growing use of NPs has raised concern over their potential harmful effects on human health

and the environment [13, 14]. This in turn has warranted thorough scientific investigations in

areas of research such as nanotoxicology. The assessment of human risks of NPs in a closed

environment with a direct exposure can sometimes be relatively straight forward when

compared with conditions in an open environment. In the former case, concentrations and

physico-chemical properties of NPs can be monitored with a defined set up [15], whereas the

latter situation is much more complex when NPs are released to the

2

environment via e.g. industrial waste or disposal and recycling of NP-products. One aspect of

this complexity is the fact that NPs undergo different transformations due to their interactions

with various kinds of organic and inorganic constituents present in the environment, which can

alter their mobility, toxicity, and particle properties [16, 17]. For example, Ag NPs are

sulfidized upon interaction with inorganic or biogenic sul de present in sub-aquatic sediments.

This changes their agglomeration behavior, surface chemistry and dissolution properties, and

therefore their fate and potential toxicity [18]. Natural organic matter (NOM), [19] present in the environment is a complex matrix of organic

molecules formed by the breakdown of terrestrial plants and by products of aquatic plants

including phytoplankton (algae, fungi, etc.). NOM is important for the fate of NPs as it

influences their agglomeration behavior by adsorbing and changing the surface properties [20]. The molecular weight and size of the NOM depend on the source of extraction. NOM has

been classified as humic (hydrophilic) and non-humic (hydrophobic) substances where the

former includes high molecular weight organic constituents. Humic substances are further

subdivided into three categories:

1. Humic acid (HA), which precipitates out of solution at pH < 2. 2. Humin, which is insoluble at all pH. 3. Fulvic acid, which is soluble at all pH.

The organic matter investigated in this study is HA, which is the main component of humic

substances. It contains aliphatic-, aromatic-, carboxylic acid, amine-, alcohol, and carbonyl-

functional groups, to name a few. Dihydroxybenzoic acid (DHBA) is another organic molecule

studied within this work together with HA, and represents smaller sized degradation products in

NOM, schematically shown in Figure 1.

Figure 1. Structure of studied isomers of DHBA and of humic acid

2.3 Risk assessment of NPs NPs have been proposed to be categorized into four groups to organize the prediction of their

fate and toxicity: carbon NMs, metal oxide and metal NPs, quantum dots, and nano-polymers

[21, 22]. Aitken et al.[23] report that not all NPs are hazardous and within a given category as

NPs differ from each other in their properties. Hence a general risk assessment protocol may not

be applicable in predicting their fate [23]. In some cases it is important to perform individual

risk assessments [24-27]. From a regulatory standpoint, there is work ongoing to adjust NPs into

the regulatory framework, e.g. in REACH - Registration, Evaluation, Authorization and

Restriction of Chemicals, implemented in the EU [28].

3

Conceptual models have been established to identify critical processes related to environmental

risks of NPs, including: (i) the form, route and mass of NMs entering the environment; (ii)

transformation, affinity for surfaces and fate in the environment; (iii) transport and geographic

distribution in the environment; (iv) bioavailability; (v) toxic responses of individual organisms

to exposure; (vi) effects on ecological structure; and (vii) changes in ecological/biogeochemical

function in the environment [29]. NP characterization has in this context been identified as a key

component for making reliable risk assessments [30]. This includes characterization of the NP suspensions using different experimental

techniques and monitoring of their dissolution/transformation [31]. Thorough characterization

improves interpretation and quality of e.g. ecotoxicity data, by for example aiding in making

proper design decisions based on NP characteristics such as size and dissolution. Hazard

identification and risk assessments can be done taking into account the above mentioned

physical and chemical properties of NPs together with toxicity result [32-35]. Production

volumes and transport of the NPs in the environment can be used in the risk assessments [2, 36].

Bioavailability is another important parameter for NPs, with one definition being “a relative

measure of that fraction of the total ambient metal that an organism takes up when encountering

or processing environmental media, summed across all possible sources, including water and

food” [37]. The bioavailable concentration is very different from the total environmental

concentration [38]. As mentioned earlier, NPs can interact with various organic and inorganic components upon

release to the environment, which sometimes results in the formation of larger clusters of NPs,

denoted agglomerates. The mobility of these agglomerates can be hindered in the environment

due to their sedimentation [39]. Mobile NPs that are dispersed into aqueous systems and remain

in the environmental medium can be transported further in the environment [40]. The mobility

of NPs is thus one an important parameter as longer residence times in solution may enable

further transport to other environmental settings of different characteristics. Dissolution is

another important process to consider while studying bioavailability [41].

3. BACKGROUND 3.1 Choice of NPs The choice of the metal NPs (Cu, Mn, and Al NPs) of the work in this thesis was made based on

differences in electrochemical and surface oxide characteristics. Cu NPs undergo relatively

rapid dissolution in cell media at near neutral pHs [42] compared to Al NPs, which have a more

protective surface oxide that efficiently hinders corrosion and dissolution at certain conditions

[43]. Al spontaneously forms a passive surface oxide in air that remains in water of neutral and

mild acidic pHs. At alkaline or very acidic conditions, it may be readily dissolved [44]. The surface oxides of Mn NPs and Cu NPs are non-passive though efficient barriers in

many conditions. As an example, the extent of metal release from in cell media follows the

sequence; Cu NPs>Mn NPs>Al NPs [45]. However, the metal release process is highly material

and solution specific. Cu NPs are e.g. widely used in electronics, metallic inks and textiles [46-48]. The environmental

concentration of engineered Cu NPs in Asian waters river water has as an example been

estimated to 0.06 mg/L, which according to Chen et al. is a high concentration for aquatic

microorganisms [49, 50]. However, the Cu NP concentration in wastewater effluents in other

investigations in central Europe has been estimated to be much lower than this concentration

(<0.1 �g/L) [33]. There is hence a large uncertainty concerning

4

environmental concentrations of Cu NPs (and of all metal NPs). Although there are studies on

effects of NOM on Cu NP dissolution and changes in chemical and physical properties [51-53],

there is still a knowledge gap concerning the interaction mechanisms between Cu NPs and

NOM. Similarly, Mn and Al NPs are used in various applications in the society, [54, 55] but

lack predictions of their environmental concentrations. Possible risks induced by the use of such

NPs cannot be ignored [54, 56]. The lack of existing knowledge about the interaction between

NOM and these metal NPs was another reason for their investigations in this thesis. 3.2 Relevant NPs characteristics from an environmental perceptive There are several important key parameters of NPs to be able to predict their environmental fate

in freshwater environments. Such parameters include e.g. their size and surface properties (e.g.

surface charge, surface oxide composition). The size is important as it influences the tendency

to form agglomerates, subsequent sedimentation and metal release [57]. Besides, biological

uptake and subsequent effects, also largely size-dependent effects have been reported, e.g. by

Ivask et al [58] and De Jong et al [59]. Smaller sized gold and silver NPs proved to be more

toxic and more easily transported in the studied mammalian bodies than corresponding larger

sizes particles.

Surface properties are other important factors to consider while assessing NP fate [60, 61].

Metal NPs spontaneously form surface oxides in contact with ambient air, with the exception of

noble metals such as gold and platinum. The characteristics of this surface oxide changes

depending on the chemical environment. Dissolution and adsorption are for instance influenced

by the properties of the surface oxide [62, 63]. For instance, the surface oxide on Al NPs,

Al2O3, is passive in nature in pure water at a near neutral pH, making the Al NPs less reactive

compared with the Cu NPs which undergo faster dissolution due to a less protective oxide. The

oxide layers of the metal NPs investigated in this thesis have been thoroughly characterized

elsewhere by means of different surface analytical and electrochemical techniques [45]. The

composition of the surface oxides of the Al, Mn and Cu NPs are schematically illustrated in

Figure 2.

Al2O3

Mn2O3/MnO2 CuO

MnO Cu2O

Figure 2. Schematic illustration of surface oxides on the metal NPs investigated in this thesis

(adapted from [44])

The surface charge is besides the surface oxide composition and the particle size very important

parameters that govern the environmental interaction of NPs [64]. Metal NPs acquire a surface

charge when exposed to an aqueous solution. The charges on the surface originate e.g. due to

protonation or deprotonation of surface hydroxyl groups. The surface charge is important as it

determines the repulsive electrostatic force between particles and influences the adsorption

characteristics of different ligands (briefly described in section 4.3) [65].

5

3.3 Selected phenomena influencing NP fate in the environment 3.3.1 Colloidal stability The classical DLVO (Derjaguin, Landau, Verwey, and Overbrook) theory for colloidal stability

can help to estimate the stability of NPs in aqueous solutions. This theory describes the colloidal

stability by considering the balance (summation) between two forces: the attractive van der

Waals forces and the repulsive electrostatic double-layer forces between particles. The van der

Waals force is always attractive between similar particles in any media and is due to interactions

between induced and/or permanent dipoles. The van der Waals force depends on the size of the

particles and the material properties such as the dielectric constant, which is expressed by the

Hamaker constant. Metals NPs are conductive and polarizable and therefore they have high

dielectric constants and refractive indices. This leads to a significantly higher Hamaker constant

for metals than for other materials. Hence, metal NPs in aqueous solution without any surface

modifications have a high tendency to agglomerate and precipitate (sediment)due to the strong

van der Waal force [66]. On the other hand, charged NPs will develop a repulsive force, due to overlap of the ion clouds,

which give rise to an increase in the osmotic pressure. Hence, NPs with the same polarity of the

surface charge will repel each other. In aqueous solutions, NPs acquire surface charges due to

dissociation of surface oxide groups and this will result in the redistribution of ions in the

aqueous solution, Figure 3. These surface groups attract oppositely charged ions (counter ions)

from the aqueous solution and ions with similar charge (co-ions) will be depleted to make a

neutral system [67]. The concentration of counter ions will decrease and the concentration of

opposite ions will increase with the distance from the surface until equilibrium is reached out in

the bulk solution. This distribution of ions surrounding the particles is called the electrical

double layer and the potential in this layer will decay exponentially with the distance. The

repulsive electrostatic double-layer force (if strong enough) between particles improves the

colloidal stability. Although DLVO theory has been used for explaining the colloidal stability, the validity of this

theory for NPs is not so clear. It is because NPs does not always possess spherical shape as

mentioned in the classic DLVO theory. Besides, as mentioned earlier, NPs possess more surface

atoms than the bulk and so dielectric properties of NPs differs from the bulk materials [68]. These dielectric properties plays an important role in DLVO force calculations [69]. In

addition, the DLVO theory only considers the electrostatic double-layer repulsion and the

attractive van der Waals forces. However, other forces such as steric repulsion cannot be

ignored, e.g. when organic matter is adsorbed to the NPs. Steric repulsion, which originates

from volume restrictions and inter-penetration effects of the adsorbed molecules, can hence also

play an important role in particle agglomeration. When two particles with adsorbed organic

molecules such as DHBA or HA are brought close to each other, the particles can repel each

other due to overlapping/compression of the adsorbed layers. This repulsion contributes to the

repulsive forces between particles. Furthermore, the attractive van der Walls force will be

lowered for particles coated with an organic layer (to some extent depending on layer

characteristics) since the Hamaker constant for the organic layer is significantly lower compared

to the metal NP. Figure 3 illustrates these different forces acting between particles in an aqueous

medium.

6

Figure 3. Schematic illustration of electrostatic-, van der Waals- and repulsive steric

forces between particles in solution 3.3.2 Metal release/ dissolution Metal release/dissolution means that metals(ions) leave the NP and migrate to the bulk solution

as illustrated in Figure 4 [70] This process is governed by controlling factors such as the surface

oxide composition and characteristics [71], the solubility of oxide constituents, electrochemical

properties [72], solution chemistry, and concentration gradients of metal ions between the

particle surface and bulk solution. The extent of metal release depends further on material- and

surface properties as well as ligand-interactions and size [73]. For instance, Zn NPs undergoes

rapid dissolution whereas titanium dioxide and cerium oxide NPs hardly dissolve in pure water

at neutral pH. The presence of different ligands in solution (e.g. counter ions and organic

matter) may either promote dissolution by ligand-induced processes, or hinder dissolution due

to the formation of a protective layer [74].

Figure 4. Illustration of dissolution/metal release processes from the

surface of metal NPs Metal release/dissolution data provides information on whether the NPs remain as particles or

becomes dissolved and released into solution and, from speciation measurements/predictions,

whether released metals form strong complexes in solutions of low bioavailability or remain as

free ions or labile complexes of high bioavailability. These aspects are thus important to

consider when assessing the fate of NPs in the environment.

7

4. EXPERIMENTAL APPROACH 4.1 Sample preparation Cu, Al and Mn NPs were investigated in this thesis. ZnO and SiO2 NPs were in some studies

investigated for comparison. Dispersions of the NPs were prepared from stock solution with particle loadings of 2.56 g/L or 1 g/L in ultrapure water using probe sonication for 15 min or 3 min [75]. After sonication, the NP dispersions were diluted to 0.1 g/L in different exposure

media. The exposure media were 1 m M sodium perchlorate, NaClO4 (Paper I) and synthetic freshwater (Paper II, ionic strength 1.2 m M) with DHBA and HA in concentrations given in

Table 1. NaClO4 was used due to its relative low impact on corrosion, compared with e.g. sodium chloride, NaCl.

Table 1. Concentrations of model NOM in synthetic freshwater, and composition of the freshwater

Organic matter Concentration in synthetic freshwater

2,3 DHBA 0.1 mM (15.41 mg/L)

1 mM (154 mg/L)

3,4 DHBA 0.1 mM (15.41 mg/L)

1 mM (154.1 mg/L)

Humic acid 2 mg/L

20 mg/L

40 mg/L

Freshwater components Molar concentration (M)

NaHCO3 7.5·10-5

KCl 7.8·10-6

CaCl2·H2O 2·10-4

MgSO4·7H2O 5·10-5

The concentrations of HA and DHBA were chosen according to the range of environmental

concentrations of humic acid.[76]

8

4.2 Characterizations of NPs 4.2.1 Size distribution As mentioned above, the NP size is an important parameter for the transformation of NPs in the

environment as it e.g. influences sedimentation and dissolution. Therefore the evolution of

particle size over time was investigated in the media described in Table 1. Changes in size of the NPs and their agglomerates in solution were investigated by means of

photon cross correlation spectroscopy (PCCS), which is based on the principles of dynamic

light scattering (DLS). A laser beam illuminates the particle suspension and the scattered light

intensity is measured with a detector, Figure 5. Observed fluctuations in the intensity of the light

are the result particle movement (diffusion) in the solution. In PCCS, two lasers are used instead

of one, as used in conventional DLS. The simultaneous irradiation of the two lasers gives rise to

two speckles that are detected in two detectors [77]. Speckles with the same scattering vector

are only taken into consideration for further analysis and multiple scattering is hence less of a

problem.

Figure 5. Schematic illustration of DLS analysis of NP dispersion The auto-correlator connected to the photodetector then compares the intensity of the scattered

light at time t with the intensity to a time later and the fluctuations are quantified by a second

order correlation function. The software uses a non-linear least square fitting algorithm to fit the

measured correlation function to solve the correlation function decay ȗ. The decay is used to

calculate the diffusion constant D, which in turn is used to calculate the particles hydrodynamic

radius (r) by using Stokes- Einstein equation, eq.1:[78]

𝐫=𝐤𝐓𝟔𝛑𝛈𝐃 (eq. 1)

Where k = Boltzmann’s constant

T= temperature in K

η = solvent viscosity

4.2.2 Zeta potential of particles in dispersion As explained earlier in section 1.4, the surface charge is an important property of NPs. In

aqueous solutions, all particles will be charged and surrounded by a layer of oppositely charged

ions, forming an inner strongly bonded layer (the Stern layer). Outside this layer, there is the

diffuse (electrical) double layer, which consists of ions loosely bound to the particles and ions

with similar charge as the particles. The distribution of these ions changes with the distance

from the surface until equilibrium is reached. When the particles move in

9

the solution, the counter ions closest to the surface will move with the particle and the rest of the

ions in the electrical double layer will be more stationary in the solution creating a plane of

shear (slipping plane). The zeta potential is the potential measured at the slipping plane, Figure

6 [79]. A higher magnitude of the zeta potential corresponds to a more highly charged surface,

which implies improved colloidal stability due to repulsion between charged surfaces of similar

polarity [79].

Figure 6. Schematic illustration of the diffuse double layer of a NP and the position of the slipping

plane where the zeta potential is determined

The zeta potential is estimated by irradiating a laser through the sample at the same time as an

electric field is applied to it. The particles then move due to their interaction with the electric

field, a phenomenon known as electrophoresis. The particle velocity, V, is determined from the

Doppler shift from the movement of the particles due to the applied electrical field, and from

this the electrophoretic mobility is calculated. The zeta potential, ȗ, is then estimated from the

electrophoretic mobility by using Henry’s equation (eq. 2).

μ=2𝜀𝜁𝑓(𝜅𝑎)3𝜂 (eq. 2)

Where μ = mobility

𝜀= dielectric constant

ζ = zeta potential k = Boltzmann’s constant

f(ka) is Henry’s function where ț[nm-1

] is the Debye parameter, and its reciprocal is the Debye

screening length of electric double layer, and a is the particle radius. ka thus is the particle

radius to the double layer screening thickness. Two approximations are frequently used to estimate the value of f(ka), and they are

schematically presented in Figure 7. For values of ka<<1, the value of f(ka) is approximated to

1. This is called the Hückel approximation. This approximation is hence valid for a situation

when the screening length of the double layer is thicker than the particle radius, and thus

represents a situation with relatively small NPs (<<100 nm) and/or an electrolyte of low ionic

strength (mM range or lower). For example, the Hückel approximation is valid for a 20 nm NP

in a solution of 1 mM ionic strength [80]. The Smoluchowski approximation assigns a

10

value of 1.5 for f(ka)and is used when the screening length of the electrical double layer is much

smaller than the particle radius (ka>>1). This hence corresponds to a situation with larger

particle sizes and/or higher ionic strengths compared with the situation when the Hückel

approximation is valid. In this study, the Smoluchowski approximation has been used for zeta

potential measurements in freshwater without NOM added, whereas the Hückel approximation

has been used to calculate their zeta potential in freshwater containing HA or DHBA, as the

particles sizes are smaller than 100 nm and the ionic strength approx. 2 mM in these media [80].

Figure 7. Schematic illustration of particle size vs. thickness of the electrochemical double layer using the Smoluchowski and the Hückel approximations, respectively

4.2.3 Dissolution/metal release The dissolution/metal release from the metal NPs was determined by means of atomic

absorption spectroscopy (AAS). AAS was chosen because it can determine metal concentrations

in solution down to a few �g/L. Dissolution in this context relates to chemical dissolution of

metals whereas metal release in addition involves electrochemical and ligand-induced

processes. The technique relies on a solution that has been atomized and that absorbs incident

electromagnetic radiation. Light is incident from a narrow line source, a hollow cathode lamp,

which emits light of specific wavelengths depending on the specific element of interest. The

incident electromagnetic radiation passes through the atoms, absorbing photons from the

incident light. The detector at the other end shows the reduction in intensity of incident light.

The value of absorbance is then used to calculate the concentration of the specific metal in the

sample by using Beer-Lambert law. Two different modes of AAS have been used in this study:

1. Flame AAS for metal concentrations in the mg/L range. 2. Graphite furnace AAS for trace concentrations of metals (�g/L).

The concentration of the measured metal is compared with the amount of light absorbed by

known standard concentrations of the same metal in order to obtain the actual metal

concentration of the sample [81]. 4.2.4 Adsorption studies of organic matter Attenuated total reflection Fourier transform infrared spectroscopy (ATR-FTIR) was used to

study the adsorption of NOM on the NPs. A benefit of this technique is that it can be

11

employed in situ and is hence able to investigate the adsorption of molecules in solution, e.g. upon NOM interactions with the NPs. ATR-FTIR measures the absorption of infrared light by the samples. An IR beam is irradiated to

an optically dense crystal with a high refractive index in a certain angle. The evanescent wave,

created due to total internal reflectance, then interacts with the sample positioned above the

crystal leading to absorption of infrared light. The evanescent wave will thus get attenuated due

to absorption of certain vibrational transitions in the samples. The recorded attenuated IR beam

is then used to generate an IR spectrum [82].

5. KEY FINDINGS (Papers I and II) 5.1 Sonication of NP suspensions largely influences the particle size and extent

of metal release (Paper I). There are several methods to disperse NPs in an aqueous solution, e.g. using sonication probes

or bath sonicators [83, 84]. It is essential to have a dispersion method that is reproducible and

well-documented protocol. It is furthermore very important to have knowledge on how it

influences the NP properties, in terms of e.g. particle size, surface oxide composition and

dissolution [85, 86]. A dispersion method has been designed by Taurozzi et al. [83] to prepare

NP dispersions in a way that could produce reproducible results in different laboratories. This

protocol emphasized that the delivered acoustic energy of the sonication procedure should be

determined and reported. Taurozzi et al. [87] Further more reported that different sonication

methods disperse NPs to different extent. Manual shaking, vortex-blending, bath sonication and probe sonication were in this study tested

for dispersing Cu-, Mn-, Al-, and ZnO NPs [75]. Probe sonication was found to be the most

efficient method to prepare dispersions of the studied NPs in terms of smallest particles size and

lowest degree of sedimentation during sonication [75]. These findings were in line with previous

investigations [88-90]. A particle loading of 2.56 g/L is common practice for stock solutions.

With this concentration of NPs, large agglomerates were observed when dispersed and sonicated

at standardized settings using a probe sonicator for the Cu-, Mn-, Al-, and ZnO NPs (7.35 W of

delivered acoustic energy). The underlying reason is mainly that such a high particle loading

results in a high collision frequency of the NPs, which in turn results in pronounced

agglomeration, Figure 8. The stock solution concentration was therefore reduced from 2.56 g/L

to 1 g/L in order to have less sedimentation and agglomeration in the sonicated stock solutions.

12

Figure 8. Schematic illustration of the influence of NP loading in the stock solution on their extent of sedimentation and agglomeration

The extent of particle de-agglomeration and sedimentation was also seen to depend on the

amount of delivered acoustic energy during sonication, and on the properties of NPs. 1 mM

NaClO4 was used for the investigations of NP agglomeration in order to use a relatively non-

corrosive electrolyte, thereby minimize the effect of metal NP corrosion. The results showed

that the agglomerate size in the Mn NP and the Al NP stock solutions that were sonicated for

approx. 3 min (182 s) in 1 mM NaClO4 were relatively larger compared to suspensions

sonicated for approx. 15 min (882 s), Figure 9. However, the Cu NPs that were sonicated for the

longer time period (882 s) did not disperse the agglomerates significantly better, with relatively

similar particle sizes as observed in 1 mM NaClO4. It has previously been suggested that

different amounts of delivered acoustic energy are needed to disperse different NPs, which is

connected to e.g. differences in effective density and surface chemistry of the NPs [91].

Figure 9. Schematic illustration of NP agglomerates formed in Cu-, Mn- and Al NP stock

solutions(1 g/L, ultrapure water) sonicated for 3 min and 15 min followed by dilution to approx. 0.

1 g/L in 1 mM NaClO4 No significant differences in the amount of released Cu in solution (Cu able to pass through a

membrane with a pore size of 20 nm) over time were observed for Cu NPs in 1 mM NaClO4

between solutions sonicated for 3 and 15 min, Figure 10. The total amount of released Cu after 24 h could be under estimated, as the amount of soluble Cu in solution was close to the theoretical saturation point (4 mg/L) [75], and the amount of Cu, possibly precipitated or

13

bonded to large complexes (< 20 nm) was not considered during analysis. For the Mn NPs on

the other hand, as can be seen in Figure 10, the longer sonication time enabled de-

agglomeration to relatively smaller sizes. This resulted in a relatively larger specific surface

area and thus more Mn release compared with release in the shorter sonication time [75].

0.25 4 24 0.25 4 24

Figure 10. Release of Cu (left) and Mn (right) in 1 mM NaClO4 (after membrane filtration with a pore

size of 20 nm) from Cu NP suspensions (0.1 g/L) probe-sonicated for

3 and 15 min, respectively As earlier mentioned, high particle loadings of the stock solutions (2.56 g/L vs. 1g/L) resulted in

increased agglomeration of the NPs. This influenced was clearly noticed for Mn NPs as shown

in Figure 11, where less Mn was released for the highest particle loading (2.56 g/L) compared

with the lower loading (1 g/L) of the stock solution. The surface area of large agglomerates is

less compared to a relatively more dispersed system, and thus the amount of released metals

becomes less. Differences between Cu NPs in stock solutions of different loadings were not

clearly visible as the release of Cu was close to its saturation point of 4 mg/L, see previous

discussions.

0.25 4 24 Figure 11. Released amount of Mn from Mn NP suspensions (1 g/L and 2.56 g/L stock solutions)

in 1 mM NaClO4 after probe-sonication for 15 min, at a particle loading of 0.1g/L

14

It is concluded that the dispersity of the NPs depends on the stock solution concentration and

the delivered acoustic energy due to sonication, together with NP properties. Metal

release/dissolution of NPs in turn depends on the dispersity of NPs and the NP characteristics

such as surface oxide characteristics. 5.2 Rapid sedimentation of NPs in the stock solution results in significantly

lower administered doses than the nominal doses of sonicated metal NP

dispersions – an effect of strong attractive van der Waals forces (Paper I). NP investigations (e.g. nanotoxicity studies) entail reproducible and well-characterized NP

dispersions that can be repeated and used in subsequent studies. The pipetted concentration, or

dose, taken from a stock solution is normally expected to be same in all experiments and equal

to the nominal concentration of the stock solution. This may however not be the case if there is

significant agglomeration in the stock solutions. The pipetted concentrations of NPs from sonicated stock solutions were investigated to quantify

the possible error in the administrated dose induced as a result of NP sedimentation during

preparation of particle suspensions. The administered (real) doses transferred from the sonicated

stock solutions were for all metal NPs found to be significantly lower, in the range of 30-70%,

than the nominal doses, Figure 12. Observed differences between the nominal and the

administered dose were independent on particle loading and sonication time (delivered acoustic

energy).

Figure 12. Schematic illustration of the difference between the nominal and the added

(administrated) concentration of NPs pipetted from a stock solution (2.56 g/L or 1 g/L) probe-

sonicated for 15 min or 3 min in ultrapure water Observed differences in doses are predominantly related to the strong van der Waals attractive

forces between the metal NPs, which result in rapid agglomeration and concomitant

sedimentation of the NPs already in the stock solutions. This is also evident from DLVO

calculations showing that the magnitude of the van der Waals attraction is dominant over the

electrostatic repulsion for all metal NPs [75]. High van der Waals attractions are related to their

high Hamaker constants, which depend on the dielectric constant and electronic structure of the

metals [92]. Effective Hamaker constants for a core-shell geometry (i.e. metal core with a

surface oxide) were calculated to investigate the effect of the surface oxide on the metal NPs in

the DLVO calculations. Even though the surface oxide thickness reduced the Hamaker constant,

the reduction was relatively small for oxide thicknesses relevant for the metal NPs

15

of this study. This confirms a dominance of the van der Waals forces over the electrostatic

forces. The trend in nominal dose between the NPs was found to be material specific (Figure 6,

Paper I). The difference was highest for the Mn NPs for which the administered dose was less

than 40% of the nominal dose. This means that each metal NP has to be investigated

individually for determining the real administered dose. Effective density differences between

metal NPs also play a role for particle sedimentation and should be considered. It is concluded that probe-sonicated NPs form large agglomerates already in ultrapure water

during the sonication step of the stock solution and undergoes rapid sedimentation. It is hence

essential to monitor the actual concentration of transferred doses in e.g. nanotoxicity studies in

order to avoid misinterpretations. 5.3 Metals are released from Cu, Al and Mn NPs already during the sonication

step and the presence of biomolecules can enhance this process (Paper I). Sonication of metal NP suspensions evidently alters the properties of the NPs, as described in

terms of size and extent of agglomeration. Even the surface oxide characteristics become

influenced by the sonication process, properties that in turn can influence the extent of metal

release from the NPs. Since metal NPs have been disclosed to dissolve and release metals

already during the sonication step, it is vital to know if, and to what extent, this occurs. If the

metal NPs dissolve to a large extent already in the stock solution, the added dose will both

comprise metal NPs and dissolved metal species. This has of course large implications on e.g.

nanotoxicity studies where NP-specific mechanisms are being investigated. The effect of

sonication on NP dissolution was investigated in which NP dispersions (Cu, Mn, Al, ZnO) were

filtered through a 20 nm-sized pore membrane directly after sonication (15 min) followed by

immediate analysis by means of AAS, Figure 13. This measured metal concentration includes

both NPs sized less than 20 nm and ionic metal species released during sonication. However, as

shown in Figure 2 in Paper I, the NPs were much larger sized than 20 nm, which means that the

fraction sized < 20 nm in Figure 13 mainly represents dissolved metal species. Figure 13. Released metals in solution without BSA (left) and with BSA (right) after membrane filtration

with a pore size < 20 nm in stock solutions (1 g/L) of NPs of Cu, Mn, Al and ZnO probe-

16

sonicated for15 min. The asterisk for Al NP indicates that the concentrations of Al could not be quantified due to low concentrations (<0.1% of added Al)

As seen in Figure 13, metals were released from the NPs already during the sonication step of

the stock solution. Observed differences in metal release between the NPs is due to differences

in surface oxide characteristics of the metal NPs (Cu, Mn, Al). Most metals were released from

the Mn NPs followed by the Cu NPs. The passive surface oxide of Al efficiently hindered the

release of Al. The ZnO NPs were chemically dissolved (no surface oxide) relatively fast. For the

Cu NPs, the total measured Cu concentration in solution was limited by saturation with

concomitant precipitation. The addition of bovine serum albumin (BSA) to the stock solutions prior to sonication for Cu

and Mn NPs further increased the release of Cu from the Cu NPs to 4% and 32% respectively as

shown in Figure 13. This is because BSA adsorbs to the NPs with a concomitant weakening of

metal-oxygen bonds due to its surface coordination. This result in a ligand exchange with the

surface oxide that either enhances the release and/or changes the surface pH. BSA can also

enhance the release of metals due to complexation of dissolved metal species in solution driving

these processes forward. BSA has been used in several toxicity studies to stabilize NPs during

sample preparation [93-97]. The effect of using stabilizing agents such as BSA on the extent of metal release/NP dissolution

should hence be taken into consideration during e.g. toxicity investigations of NPs. Results from

such toxic assessments can otherwise be misleading as the reported toxicity could be due to

released metal ions rather than metal NPs. This is e.g. important since metal NPs and released

ions and formed complexes have different affinities and abilities to interact with cell membranes

[98, 99]. The presence of adsorbed BSA, a bio-corona, will also influence subsequent cell-

interactions. 5.4 The extent of particle agglomeration of the Cu, Al and Mn NPs depends on

the type of surface adsorbed NOM (Paper II).

If NPs are dispersed into the environment, their interaction with NOM is one key factor that will

influence their agglomeration behavior since, if adsorbed; it will change the surface properties

of the NPs. It is hence important to study the interaction between NPs and NOM, including

different kinds of NOM, in order to enable predictions and improve the current understanding

on how metal NPs will change their size once dispersed in e.g. in a freshwater setting.

Figure 14 shows the size distributions of Cu, Mn, and Al NPs approx. 15 min after being

sonicated in ultrapure water and diluted to freshwater containing 1 mM 2,3 DHBA, 1 mM 3,4

DHBA, or 20 mg/L HA. It is evident that both DHBA and HA in freshwater, at least initially,

counteracted agglomeration compared with freshwater only. The NPs agglomerated and

sedimented in less than 30 min upon exposure in freshwater without any organic component.

This is, as discussed in section 5.3, predominantly due to high van der Waals forces between the

metal NPs, and manifested by the large agglomerate sizes observed in Figure 14.

17

Figure 14. Size distributions of Cu, Mn and Al NPs in synthetic freshwater with and without NOM

(1 mM 2,3 DHBA, 3,4 DHBA or 20 mg/L HA), measured approx. 15 min after solution sonication The results show that HA in a concentration of 20 mg/L kept the NP agglomerate size close to

the primary sizes of the Cu and Al NPs (see Paper II). This reduction in agglomerate size upon

the addition of HA, also observed for the DHBA monomers, Figure 14, is related to the

electrostatic and steric repulsion induced by adsorbed HA or DHBA. Surface adsorption of

DHBA or HA to the NPs was confirmed by the ATR-FTIR measurements (Figure 2, Paper II).

The observed size distributions of the NPs in freshwater containing NOM depended on its

concentration in freshwater (Figure 7, Paper II). For example, size distributions of the different

metal NPs are presented in Figure 15 for different concentrations of HA.

Figure 15. Relative differences in size of metal NP agglomerates in synthetic freshwater only

compared with freshwater containing HA of different concentration (pH 6.2) The size of NP agglomerates formed in freshwater containing 20 mg/L HA was the smallest

among all three concentrations of HA. Cu and Al NPs formed relatively large agglomerates in 2

mg/L and 40 mg/L HA compared to 20 mg/L HA, as also shown in Figure 7, Paper II. This may

be connected to the ability of HA to cross-link at high concentrations (40 mg/L) with increased

agglomeration as a result, see Ghosh et al.[100]. At low HA concentrations (2 mg/L), the steric

repulsion may be insufficient (too thin layer of adsorbed HA) to overcome and screen the van

der Waals forces. Observed changes in surface charge were expected based on the adsorption of DHBA and HA.

Differences in surface charge, represented in terms of the measured zeta potential, are shown for

the different metal NPs and NOM in Figure 16.

18

Figure 16. Differences in zeta potential of the metal NPs (0.1 g/L) in synthetic freshwater (pH 6.2)

only and when containing either 1 mM 2,3 DHBA, 3,4 DHBA or 20 mg/L HA) measured approx.

15 min after the sonication The zeta potential became more negative in freshwater in the presence of NOM, which indicates

adsorption of NOM to the NPs. The magnitude of the zeta potential depended on the type of

NOM and also on the orientation of functional groups linked to the particle surface. For

example, the Mn NPs were more charged in freshwater containing 1 mM 2,3 DHBA compared

with freshwater with 1 mM 3,4 DHBA. All NPs became more charged, and hence more stable

in solution in the presence of HA compared with DHBA (briefly discussed in section 5.4).

The size of the agglomerates depends on the concentration of DHBA (Figure 7, Paper II). For

example, agglomerates of Cu NPs in freshwater containing 0.1 mM 2,3 DHBA or 0.1 mM 3,4

DHBA were larger compared to corresponding sizes in freshwater containing 1 mM of the

DHBA monomer. In all, it is evident that the understanding of the fate of metal NPs in the environment requires

knowledge also on concentrations and types (e.g. size, functional groups) of NOM and on its

influence on other phenomena including e.g. dissolution, agglomeration, and sedimentation of

metal NPs, as discussed in sections 5.4 -5.6. 5.5 NOM delayed the sedimentation of the Cu and Al NPs, in particular

pronounced in the presence of HA, whereas no significant effect was observed

for the Mn NPs (Paper II). NOM is present in any natural water and is a complex mixture of different organic molecules,

and its composition and properties varies depending on different geographical areas. This

variation gives it e.g. altered properties in terms of metal complexation. It is hence important to

look at different kinds of NOM when assessing its influence on the metal NP stability. NP size distributions were monitored for 24 h in freshwater containing 20 mg/L HA to study

agglomeration and sedimentation. As described in section 5.3 above, HA in concentration 20

mg/L was found to disperse NPs close to their primary size. The dispersity of the NP

suspensions in 20 mg/L HA was retained up to 24 h as shown in Figure 17. However, for the

higher HA concentration (40 mg/L) rapid sedimentation of the NPs took place within 4 h. At the

lowest concentration of HA investigated (2 mg/L), the Cu and Al NPs underwent complete

sedimentation within 6 h whereas the Mn NPs tended to be more stable with much lower

sedimentation velocities.

19

Figure 17. Size distributions of NPs in synthetic freshwater with 20 mg/L humic acid over 24 h In DHBA, Cu and Al NPs started to sediment after 4 h with complete sedimentation after 6 h.

Metal NPs in freshwater of lower (0.1 mM) DHBA concentrations sedimented faster compared

with corresponding conditions with the higher DHBA concentration (1 mM). This implies that

even small organic molecules like DHBA postpone the agglomeration of these metal NPs when

compared to freshwater only. In contrast, the Mn NPs remained stable in the presence of both

DHBA and in HA up to 24 h. The sedimentation velocity of the metal NPs was calculated by taking into account the time of complete sedimentation of the NPs in fresh water with and without NOM (Table 4, Paper II). These numbers can be used in fate modelling to predict their environmental mobility. In general, the numbers imply limited mobility due to rapid sedimentation, with the exception of conditions with HA when sedimentation in many cases was significantly delayed. In contrast to the metal

NPs, SiO2 NPs were not influenced by either the presence of DHBA or HA and readily

remained stable up to 21 days (Figure 9, Paper II). One factor contributing to this may be due to

its high charge of SiO2 together with significantly lower Hamaker constant compared with the

metal NPs and hence smaller attractive forces between the NPs [101]. The dependence of the type and concentration of NOM on NP sedimentation was strong for the

metal NPs. Smaller sized DHBA provided less protection against agglomeration, showing that

it, in this context, cannot function as a model for the full, complex NOM, represented by HA.

5.6 The release of metals from the Cu, and Al NPs was enhanced in the presence

of NOM, whereas no significant effect was observed for the Mn NPs despite NOM

adsorption. Released Cu and Al readily formed strong NOM-complexes of low

bioavailability in solution whereas released Mn remained as free ions (Paper II).

The interaction between NOM and metal NPs also influenced the extent of particle

dissolution/metal release from the NPs. This is a complex interplay between many different

phenomena including corrosion, surface oxide-NOM interactions, solution chemistry, and

changes in surface pH, to name a few aspects. Dissolution/release rates can be used for

environmental fate modelling of NPs in different aquatic settings and be used to improve the

mechanistic understanding of how NOM, e.g. due to surface complexation or de-agglomeration,

influences the propensity of metal NPs to release metals.

20

As an example, the extent of Cu release from the Cu NPs was enhanced in freshwater

containing DHBA or HA compared with fresh water only. This is schematically illustrated in

Figure 18. The same situation was observed for the Al NPs. In contrast, no large effects of

NOM were observed for the Mn NPs, and scarce complexation with NOM was predicted for

released Mn ions, remaining mainly as free ions in solution. Further details are given in Figure

10, Paper II.

Figure 18. Schematic illustration of the release of Cu from Cu NPs in synthetic freshwater only

and in freshwater with either the DHBA monomer or HA. Speciation predictions of released Cu in

solution are illustrated This enhancement in metal release can partially be due to surface complexation of the DHBA

and HA molecules to the surface oxide of the NPs as seen using ATR-FTIR (Paper II),

conditions that weaken the metal-oxygen bonds and thus increase the extent of dissolution/metal

release [102, 103]. Corrosion of NPs due to a reduced surface pH upon adsorption of NOM

could be another possible reason for an increased amount of released metals/particle dissolution.

This will also influence the metal concentration in solution, seen especially for Al and Cu that

both have relatively low solubility in freshwater. More Cu in solution was released from the Cu

NPs into freshwater containing the 2,3 DHBA molecule than if containing the 3,4 DHBA

molecule. This could be related to the formation of a binuclear complex with the 3,4 DHBA

molecule that decelerates dissolution compared to the 2,3 DHBA molecule (Paper II) with

which Cu forms mononuclear complexes.[104] The Cu solubility was higher in freshwater

containing 3,4 DHBA compared with 2,3 DHBA, which means that the enhanced extent of

released metals into solution observed with the former is not because of a relatively higher

solubility of Cu in this medium. Conversely, the release of Mn from the Mn NPs was not

significantly enhanced by the introduction of DHBA (Figure 10, Paper II), although DHBA did

adsorb to the Mn NP surface, as deduced by ATR-FTIR (Figure 2, Paper II). Similar

observations have previously been reported where Mn release was seen to be unaffected by cell

media containing proteins [42]. This concludes that the adsorption of organic molecules to the

surface is probably not the rate limiting step in the dissolution process of the Mn NPs of this

study. In all, the results show that the dissolution/metal release rates of the metal NPs were relatively

high in freshwater in the presence of DHBA or HA, with 50% of the particle mass estimated to

be dissolved within 7-60 h, based on a first order dissolution rate equation (see Paper II). This

implies that the influence of NOM will reduce the mobility of these metal NPs in freshwater

settings and, that a large fraction of the metal NPs will dissolve. This implies that released Cu

and Al from the Cu and Al NPs predominantly will form non-bioavailable

21

complexes in solution. Released Mn will not form complexes to the same extent and most

probably mainly remain as free ions.

6. MAIN MESSAGES

It was concluded that sonication plays a vital role in size and dissolution of NPs as described in

Section 5.1-5.3 and schematically illustrated in Figure 19. Sonication parameters thus should be

standardized for individual NPs depending on their surface characteristics and kind of intended

investigation. In addition, characterization of surface properties and the administered dose of

NPs (transferred from stock solution) are very important aspects when investigating effects of

metal NPs.

Figure 19. Schematic illustration of the effect of sonication on the

dispersion of NPs metal NPs in suspension

The study further illustrates that NOM, such as DHBA and HA, adsorbs to the NPs and

increases their colloidal stability by electrostatic and steric repulsion, though the strong van der

Waals forces of metal NPs dominate (Sections 5.4-5.6). However, the extent of de-

agglomeration and stability depends on the type of NOM and its concentration in freshwater as

shown schematically in Figure 20.

22

Figure 20. Schematic illustration of the effect of NOM in freshwater on the extent of

agglomeration, sedimentation, dissolution/metal release and speciation for metal NPs Even though all concentrations and types of NOM to different extent enhanced the particle

stability of the NPs of this study, in particular for synthetic freshwater containing 20 mg/L HA,

the results imply a possibility for an increased mobility whereas an increased NP dissolution in

the presence of NOM will limit their potential transport.

7. FUTURE WORK

A few proposals on future research directions related to this work.

1. Detailed study on NP-NOM interactions. Further detailed investigations of NP-NOM

interaction and their chemistry will provide knowledge about their properties such as

stability and dissolution in environment.

2. Effect of sonication on organic matter stabilization. It would be relevant to study the

agglomerate sizes and stability of NPs when they are exposed to organic molecules

without sonication. In the actual environments, where particles will get exposed without

sonication, it would be interesting to assess the concentration of organic matter required

to stabilize NPs.

3. Behavior of NPs in different compartments of the aqueous environment. As pH and

ionic strength vary in different aqueous settings in the environment, it would be

interesting and relevant to study the effect of these conditions on NP stability and

mobility.

4. Combinatory effects of fulvic and humic acid. Besides HA, Fulvic acid (FA) is the

other important fraction of humic substances present in environment. It has relatively

lower molecular mass than HA but contains more hydrophobic groups such as alkanes

and aromatic hydrocarbons. Since HA contains more hydrophilic groups (hydroxyl and

carboxyl), it is easier for HA to transfer from the water phase to the surface of

23

NPs [105]. This leads to the assumption that FA may result in less stable NPs aggregates

than HA, as observed for TiO2[106]. However for some NPs such as Ag, FA tends to

stabilize the particles more than HA. Interactions between FA and metal NPs would thus

be interesting and relevant to look into in the context of NP risk assessment.

8. ACKNOWLEDGMENTS

I would like to thank everyone who contributed in some way to the work described in this

licentiate thesis. It has been a learning experience for me on both a scientific and a personal

level. First and foremost, I am thankful for the funding provided by Mistra. I feel lucky to be a part of

Mistra programme- Mistra environmental nanosafety. Many thanks and appreciation for all my

supervisors for their support, motivation, patience and continuous optimism. This work would

hardly have been completed without their guidance. My deep gratitude to Inger Odnevall

Wallinder for giving me the opportunity to perform in this project. I would like to thank her for

the guidance, useful critiques, advice and assistance in keeping my progress on schedule which

have always push me positively. Many thanks to Susanna Wold for countless interesting

discussions on the project and other fun chatting. I would like to thank Eva Blomberg for

sharing her expertise in DLVO theory and colloidal chemistry. I would particularly single out

Jonas Hedberg for all the guidance and cooperation during these two years. Special thanks to

Jörgen Rosenquist for all the discussion while in Göteborg and patience to answer my queries

through long emails. Many thanks to all the co-authors from Karolinska Institutet in the

publication included in this thesis for the valuable insight to this project.

I would like to thank Yolanda Hedberg for the help during the initial days in the group. I always

look you as a best role model for a scientist, mentor and teacher. Many thanks to Gunilla

Herting for all discussions about AAS and interesting lunch room discussion. I would also like

to take the opportunity to thank Mark Rutland to review my thesis during his summer vacation.

I thank master students Frederik Mathiason, Sara Isaksson and Maria-Elisa Karlsson (Shorty)

for the help in my experiments during sunny summer in Sweden. Ex PhD student and a dear

friend Sara Skoglund for teaching me the use of our very dear instrument - PCCS beside

continuous support and encouragement. Thank you to Eva Luna at RISE for helping me in

Zetasizer measurements at any time of my need. Thank you to Nazanin Alipour, PhD student at

the polymer department for the cooperation on the sonicator standardization. I would like to thank all my former and present friends at surface and corrosion division for

listening, offering me advices and supporting me through the entire process. Special thanks to

Laetita (Thank you for the help during thesis writing), Adrian, Cem, Laetita, Krishnan,

Akansha, Tingru, Gen, Min, Marie, Fan, Erik, Georgia, Angelica, Neda, Golrokh, Zahra for

lighting up the situation to a fun mode every day. Last but not the least, my gratitude to my three member family in Sweden and the big fat family

back in India for their endless love, support and patience so that I could dedicate precious family

time and focus in the development of my research interest.

24

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