Materials and methods - Ghent...

GHENT UNIVERSITY

FACULTY OF BIOSCIENCE ENGINEERING

CENTER FOR ENVIRONMENTAL SANITATION

Academic Year 2011 – 2012

PHYSICOCHEMICAL FATE OF ENGINEERED METALLIC NANOPARTICLES IN AQUATIC ENVIRONMENTS

Tewodros Tilahun Geremew

Promoter: Prof. dr. ir. Gijs Du Laing

Tutor: M.Sc. Frederik Van Koetsem

Master’s dissertation submitted in partial fulfillment of the requirements for the degree of

Master of Science in Environmental Sanitation

i

GHENT

UNIVERSITY

FACULTY OF BIOSCIENCE ENGINEERING

CENTER FOR ENVIRONMENTAL SANITATION

Academic Year 2011 – 2012

PHYSICOCHEMICAL FATE OF ENGINEERED METALLIC

NANOPARTICLES IN AQUATIC ENVIRONMENTS


Promoter: Prof. dr. ir. Gijs Du Laing

Tutor: M.Sc. Frederik Van Koetsem

Master’s dissertation submitted in partial fulfillment of the requirements for the degree of

Master of Science in Environmental Sanitation

ii

COPYRIGHT

The author, the promoter and the tutor give permission to use this thesis for consultation and to

copy parts of it for personal use. Any other use is subject to the Laws of Copyright. Permission

to produce any material contained in this work should be obtained from the author.

© Gent University, August 2012

The Promoter

Prof. dr. ir. Gijs Du Laing

The Tutor

Frederik Van Koetsem

The Author


iii

ACKNOWLEDGEMENTS

Glory to God

I would like to thank Prof. dr. ir. Gijs Du Laing, my promoter, for offering me the chance as a

master student to work in laboratory of Analytical Chemistry and Applied Ecochemistry. I have to

appreciate him for his valuable comments too.

I wish to express my deepest appreciation to Frederik Van Koetsem (PhD candidate), my tutor, for

his valuable comments and excellent supervision. I must also thank him for the cordial relations

showed towards me, which was very helpful and is very much appreciated.

Thanks to Joachim, Martin, Ria, Katty, and other staff from the Lab of Analytical Chemistry and

Applied Ecochemistry, Ghent University, for their unreserved support during my laboratory thesis

work.

I am extremely thankful towards the Flemish Interuniversity Council, Vlaamse Interuniversitaire

Raad (VLIR), for the generous scholarship that helped me pursue my studies at Ghent University,

Belgium. I am grateful to the people working in the Belgium Embassy in Ethiopia, especially Lea

Feleke for her unreserved help.

I would also like to extend my gratitude to CES program promoter, Prof. Marc Van den Heede, for

letting me follow this master program. I am very thankful to Center for Environmental Sanitation

program coordinators: Sylvie Bauwens, Veerle Lambert, and Isabel Depotter, for their concern

and dedication to help students and for the collegial atmosphere that I cherished during my stay at

CES.

3A’s thank you a lot, you taught me how to think but I …. Thank you 3A’s + A for your kindness

and whatever good that you did for me. Thank you AY For your good wishes.

Thank you Kessis Dr. Argaw Ambelu for all what you did, I am so happy for that.

Prof. Tefera Belachew and Dr. Fantahun Wassie thank you very much for your unreserved help.

Not to forget Jimmy, Mari Stella Park, and Jihoon, thank you for your pray to my success and the

gift that you and I both know. Jihoon thank you my sister.

iv

ABSTRACT

Engineered nanoparticles (ENPs) are particles intentionally produced by human for different

purposes. Their fate and behavior in different aquatic environments such as rivers, lakes,

groundwater, or seawater is not sufficiently understood. Moreover, the introduction of

nanoparticles (NPs) in aquatic environments will likely cause toxicity for aquatic organisms,

influence microstructures, pathogen chemistry, bioavailability, transport of contaminants and

nutrients. In order to understand the probable behavior and fate of NPs in the aquatic

environment, it is necessary to understand their interaction with natural water components in

different physicochemical conditions.

Therefore, batch experiments were setup to examine partitioning behavior of CeO2 and Ag

engineered nanoparticles in suspensions of sediments sampled from different locations in the

River Scheldt, differing in physicochemical characteristics. CeO2 and Ag engineered

nanoparticles as well as their corresponding Ce (III) and Ag (I) ions were spiked into the

sediment suspensions. Cerium and Ag concentrations were analysed in the supernatant after

centrifugation of samples taken at different equilibration times. Prior to this partitioning

experiment, experiments were conducted to assess the impact of centrifugation speed on

suspended matter remaining in suspension and to test the effect of different filtration procedures

during sample preparation.

Background concentrations of both Ce and Ag in the supernatant of sediment suspensions were

under the detection limit. The concentration of Ce and Ag in the supernatant previously spiked

with either ENPs or ions significantly depended on equilibration time and centrifugation speed.

Remarkably, CeO2 and Ag ENPs were found to be more mobile, i.e. more present in the

supernatant, than their corresponding Ce (III) and Ag (I) ions. The Ce and Ag concentrations

observed in supernatant differed significantly between different sediment suspensions,

suggesting that sediment properties influence the partitioning behavior of the ENPs. CeO2

nanoparticles were found to be more mobile in the suspensions of two sediments compared to

two other sediments, and compared to Ag nanoparticles. This may possibly be due to the higher

v

pH, and higher chloride and carbonate contents in these sediments. When spiking CeO2

nanoparticles, Ce is mobilised to the supernatant during the first 2 hours after which it is

immobilised again. This was also the case for Ag in the one sediment, whereas mobility of Ag

continuously decreased in the other sediments. The latter may have been due to a rapid release of

Ag+ ions from the Ag nanoparticles and association of the released Ag+ ions to particulate

material in suspension.

Key words: Nanotechnology, Nanoparticles, Environmental fate, Aquatic environment, Metals,

Sediments.

vi

TABLE OF CONTENTS

Contents Page

Acknowledgements iii Abstract iv Table of contents vi List of tables viii List of figures ix List of abbreviations xi

1. INTRODUCTION 1.1. Background 1 1.2. Statement of the problem 1 1.3. Significance of the study 2 1.4. Objective of the study 3 2. LITERATURE REVIEW 2.1 Nanoparticles 4 2.2 Types, sources, and applications of nanoparticles 4 2.3 Nanoparticles and colloids in aquatic environments 6 2.4 Processes affecting the environmental fate of engineered nanoparticles (ENPs) 8 2.4.1 Aggregation 8 2.4.2 Deposition 10 2.4.3 Solubility and dissolution 10 2.5 Nanoparticle and contaminant transport 11 2.6 Reactivity of nanoparticles 12 2.7 Ecotoxicity and metallic nanoparticles 13 2.8 Silver nanoparticles and silver ion 15 2.9 Cerium dioxide nanoparticles 15 2.10 Other most common metallic ENPs 2.10.1 Titanium dioxide nanoparticles 16 2.10.2 Iron nanoparticles 17 2.10.3 Gold nanoparticles 18 2.10.4 Tin dioxide nanoparticle 18 2.11 Sediment 19

vii

3. MATERIALS AND METHODS 3.1 Introduction 20 3.2 Sediment sampling and characterization analysis 20 3.2.1 Sediment sampling and preparation 20 3.2.2 Sediment characterization 21 3.3 Test to assess impact of centrifugation speed 26 3.4 Effect of centrifugation speed on TOC concentration 27 3.5 Partitioning behavior of metallic ENPs in sediment suspension 27 3.5.1 Screening experiments 27 3.5.2 Partitioning experiment 29 4. RESULTS 4.1 Sediment characteristics 31 4.2 Assessment of impact of centrifuge speed 34 4.3 Assessment of the effect of centrifuge speed on total organic carbon in SN 35 4.4 Screening experiment 35 4.5 Partitioning experiment 40 5 DISCUSSION 5.1 Sediment characteristics 46 5.2 Assessment of impact of centrifuge speed 48 5.3 Screening experiment 49 5.4 Partitioning experiment 51 6 CONCLUSIONS AND RESEARCH PERSPECTIVES 6.1 Conclusions 58 6.2 Research perspectives 59 References 60

viii

LIST OF TABLES

Table

No.

(Table) Title Page

2.1 Some examples of NPs source and application. 5

3.1 Content in the centrifuge tubes during the partitioning experiments. 30

4.1 Physicochemical characteristics of the River Scheldt sediments. 31

4.2 Metal and trace element content in the sample sediments determined via ICP-

OES after aqua regia digestion.

32

5.1 Averages of some of trace metal concentration observed in sediments of our

study and reference values for comparison.

48

5.2 Mean concentration of spiked sediment suspension showing statistical description with respect to time and centrifugation speed.

55

5.3 Pearson’s correlation coefficient between Ce and Ag concentration in SN of

sediment suspension spiked with Ag/Ce ions or Ag/CeO2 ENPs and sediment

characteristics.

57

ix

LIST OF FIGURES Figure

No.

(Figure) Title Page

2.1 Classification of NPs in the environment. 6

2.2 Size domain and typical representatives of natural colloids and nanoparticles. 8

2.3 Generalized trend for size dependent reactivity change of a material as the particle

transitions from macroscopic (bulk-like) to atomic.

12

3.1 Sediment sampling site. 20

3.2 Experimental flow sheet of screening experiments. 28

3.3 Centrifuge tube used for partitioning experiments. 29

3.4 Experimental procedure for partitioning experiments 30

4.1 Texture analysis results for the sediments sampled at Doel, Linkeroever, Bornem and

Mariekerke.

33

4.2 Amount of suspended matter in function of centrifuge speed. 34

4.3 TOC content in the SN of the different sediment suspensions in function of centrifuge

speed.

35

4.4 Cerium concentration in suspension of blank sediment, control ions and control ENPs

(without sediment) and sediment suspensions spiked with ions or CeO2 ENPs.

36

4.5 Cerium concentration in aqua regia digests of suspensions obtained after gravity settling

or centrifuging Milli-Q water previously spiked with Ce ions and Ce ENPs at different

speeds.

37

4.6 Cerium concentration in the SN of sediment suspension previously spiked with Ce ions

and Ce ENPs.

38

4.7 Cerium concentration in filtrates obtained after microfiltration of SN collected after

centrifugation at different speeds.

39

4.8 Cerium concentration in total solution and filtrate of blank, controls and spiked

sediments after different filtration steps.

39

4.9 Cerium concentration in aqua regia digests of total solution of blank and spiked

sediment suspension.

40

4.10 Cerium and Ag concentration in aqua regia digests of total solution. 42

x

4.11 Cerium concentration in supernatant after different centrifuge speed. 43

4.12 Silver concentration in supernatant after different centrifuge speed. 44

5.1 Chloride content of River Scheldt in relation to distance from the mouth of the river 47

5.2 Classification of the sediment of Doel, Linkeroever, Bornem, and Mariekerke using a

texture triangle.

48

5.3 Amount of suspended matter staying in the SN after centrifugation. 49

5.4 Concentration of Ce and Ag in the supernatant after gravitational settling of ENPs

spiked sediment suspension.

54

xi

LIST OF ABBREVIATIONS

AR Aqua regia Ag+ Silver ion Au3+ Gold ion CEC Cation exchange capacity CeO2 Cerium dioxide ICP-OES Inductively coupled plasma-optical emission spectrometry Ctr ion SN Control ion in supernatant Ctr NP SN Control nanoparticle in supernatant Ctr ion T Control ion in total solution Ctr NP T Control nanoparticl in total solution DLS Dynamic light scattering DM Dry matter DOC Dissolved organic carbon EC Electrical condactivity ENPs Engineered nanoparticles Fe2O3 Iron (III) oxide ICP-MS Inductively coupled plasma-mass spectrometry OM Organic matter NOM Natural organic matter NPs Nanoparticles ROS Reactive oxygen species SN Supernatant SnO2 Tin dioxide TiO2 Titanium dioxide TN Total nitrogen TOC Total organic carbon TP Total phosphorous t0 Time at 10 min t2 Time at 2 hours t24 Time at 24 hours

Introduction

1

1. INTRODUCTION

1.1. Background

Nanotechnology involves the synthesis, manipulation, assembly and application of materials in

the range of 1 to 100 nm at least in one of the three dimensions (Fabrega et al., 2011). Currently

the nanotechnology area has achieved a billion US dollar market and it is expected to grow to 1

trillion US dollars by 2015 (Aitken et al., 2006). Its rapid development has introduced

nanoparticles (NPs) into many aspects of our daily life. For example, Engineered Nanoparticles

(ENPs) are utilized in a variety of areas such as pharmaceuticals, cosmetics, electronics, optical

devices, environmental remediation, catalysis chemistry and material sciences (Ju-Nam and

Lead, 2008). These ENPs can be metal, metal oxide or carbon based. They may or may not have

additional surface coating. Among ENPs metallic engineered nanoparticles (metallic ENPs) have

received much attention and are now being used in different application areas (Miao et al., 2010).

The increasing use of nanotechnology in many industrial processes and consumer products will

inevitably lead to the release of nanoparticles and products containing them into the natural

environment during the product's life cycle.

1.2. Statement of the problem.

The small size gives ENPs many interesting properties like rapid diffusion, high specific surface

areas, and reactivity in liquid or gas phase (Thill et al., 2006). However, these properties may

also have unwanted effects. For instance, metal oxide nanoparticles like TiO2 and Fe2O3 can

enter into the human body and cause some toxicity, such as a cytotoxicity response,

inflammatory response, and cell membrane leakage (Brunner et al., 2006). In addition, it was

reported that silver engineered nanoparticles (Ag ENPs) are toxic to marine phytoplankton (Miao

et al., 2009) and the review of Borm et al. (2006 b) indicated that ENPs may also induce toxicity

in animals. Moreover, ENPs can enter the ecosystem and accumulate in air, water, soil, or

organisms from point sources such as factories or landfills as well as from nonpoint sources such

as through wet deposition, runoff, and abrasion from products containing ENPs (Wisner et al.,

2006). They might be highly mobile and quickly transported in the environment, or inside the

human or animal body via water or air (Howard and Wim, 2004).

Introduction

2

Although there are indications that ENPs may be transported in the environment and exhibit

toxicity towards animals, humans and plants, the current knowledge on the fate of ENPs is not

sufficient and thus, the rapid development of ENPs raises questions about their impact on the

environment, plant, animal and human health. Furthermore, their specific properties lead to new

means of interactions with environmental systems which can have unexpected impacts. Even

though their impact is not sufficiently known, it is clear that the behaviour of ENPs in the

environment, their distribution, uptake, and effects within living organisms is likely to be

different when compared to the bulk material and other xenobiotics (Scown et al., 2010).

In general, wind or runoff can transport ENPs into aquatic systems from direct discharges,

accidental spillages, wastewater effluents or solid waste dump sites. Environmental releases from

spillages during transportation of ENPs from one site to the other, intentional release when used

as catalyser during in situ environmental remediation as well as diffuse release during wear and

erosion from general use enhance environmental exposure of ENPs.

1.3. Significance of the study

The transport of ENPs in the environment is inseparable from the risk prediction ( Handy et al.,

2008) because once released into the environment they either stay at the point of release or are

transported away and are redistributed in the environment. If transport is easy, the materials may

distribute widely in the environment, and local concentrations would be relatively low. If they do

not readily migrate from the point of release, local concentrations may be high. Thus, in order to

predict the risks ENPs may pose it is necessary to have a better understanding of the factors

controlling their transport in the environment. Therefore, this investigation was designed to

contribute to environmental exposure awareness of ENPs through identification of factors

determining fate and transport of metal(lic) ENPs in aquatic environments.

Introduction

3

1.4. Objective of the study

The general objective of this study is to contribute to environmental fate studies of metallic

engineered nanoparticles released into surface water, and reveal kinetics of partitioning processes

occurring after their release into the aquatic environment. In particular, this study aims to: (1)

describe the partitioning of silver (Ag) and cerium dioxide (CeO2) ENPs between the aquatic

phase and sediments after their release into sediment suspensions, and (2) identify factors

affecting the partitioning, fate and transport of Ag and CeO2 ENPs in surface waters in contact

with sediments.

Literature review

4

2. LITERATURE REVIEW

2.1. Nanoparticles

Nanoparticles (NPs) are particulate matter with at least one dimension smaller than 100 nm. This

definition categorizes NPs in a similar size domain as that of ultrafine particles and considers

them as a sub-set of colloidal particles (Warheit, 2008; SCENIHR, 2007). They can be tubular,

spherical, or irregularly shaped and can exist in agglomerated forms (Nowack et al., 2007;

Aitken et al., 2006). They are characterised by physical parameters, such as size, shape, surface

area, molecular weight, and chemical composition. For instance, 35-40% of the atoms are

localized at the surface of a 10 nm particle compared to only 20% for 30 nm particles. Such

elevated surface area to volume ratio may be associated with multiple intrinsic properties that are

size dependent, like strong surface reactivity (Auffan et al., 2008). NPs, particularly those

smaller than 20 nm, may exhibit new properties when compared with bulk materials. Gold

particles for example are chemically inert and resistant to oxidation when they are at

macroscopic scale. However, they have a chemically active surface and may be utilized as

catalysts at nanometre size (Wang and Ro, 2006; Chiang et al., 2006). These new size dependent

properties make NPs desirable for technical and commercial use. In contrast, they may create

concerns in terms of toxicological or environmental impact.

2.2. Types, sources and applications of NP

In general NPs can be classified as naturally occurring and anthropogenic. Based on their

chemical composition they can be categorized into carbon containing and inorganic (mostly

based on metal and metal oxide) NPs. Naturally occurring carbon containing NPs can be of

biogenic, geogenic, atmospheric or pyrogenic origin, whereas anthropogenic NPs are either

unintentionally, as by-product, or intentionally produced. Intentionally produced NPs are often

referred to as engineered nanoparticles (ENPs), such as Ag and CeO2 NPs (Nowack et al., 2007).

Literature review

5

Due to their unique properties these NPs play an important role in nanotechnology progress

particularly in the following areas: electronics, food and food packaging, cosmetics, paints,

coatings, pharmaceuticals, biomedicine, catalysis and material science, environmental analysis

and remediation (Silva et al., 2011; Ju-Nam and Lead, 2008; Aitken et al., 2006).

Some examples of ENPs sources and applications are given in Table 2.1 (Brar et al., 2010;

Weinberg et al., 2011).

Table 2.1 Source, types and applications of ENPs.

Source Types Application

Metals and alkaline earth metals

Ag Antimicrobials, paints, coatings, medical use, food packaging, socks, textiles, wound dressings

Fe Water treatment, detoxification of organochlorine pesticides

Au Electronics in flexible conducting inks or films, and as catalyst, in air filters, toothpastes, baby products, vacuum cleaners, and washing machines, as vector in tumour therapy

Sn Paints

Metal oxides

TiO2 Cosmetics, paints, coatings, sunscreen lotions

CeO2 Fuel catalyst

Source: Brar et al., 2008; Weinberg et al., 2011.

The classification of NPs in the environment as described by Bhatt and Tripathi (Bhatt and

Tripathi, 2011) is presented in Figure 2.1.

Literature review

6

Figure 2.1 Classification of NPs in the environment (Bhatt and Tripathi, 2011).

2.3. Nanoparticles and colloids in aquatic environments

Release of washing powders into municipal wastewater systems, domestic washings of

cosmetics, accidental release from industries through rinsing or during transportation, and using

NPs in soil or water cleansing are some of the pathways releasing NPs into aquatic environments

(Cumberland and Lead, 2008). The major physicochemical pathways that control the fate of NPs

in these aquatic environments are: dissolution, adsorption to particulates and other solid surfaces,

aggregation and sedimentation, binding to natural dissolved organic matter, stabilisation by

surfactants, biological degradation, and abiotic degradation like hydrolysis and photolysis, as

well as oxidation and reduction in some environments for specific particles (Batley and

McLaughlin, 2010). Among these pathways, aggregation and dissolution are the most important

contributors to the environmental impacts of ENPs in water (Batley and McLaughlin, 2010).

Literature review

7

Nanoparticles can be present as suspension (solid in liquids) or as emulsion (two liquid phases)

(Brar et al., 2010). They can form a colloidal suspension in water and may interact with each

other or with other colloidal material since colloids show Brownian motion and interaction with

other particles and dissolved molecules. They may be transported easily from water to sediment

by forming aggregates and agglomerates that likely precipitate (Velzeboer, 2008). However, the

surface and interfacial properties of NPs can be modified with chemical agents like surfactants.

These agents can indirectly stabilize NPs against coagulation or aggregation by conserving

particle charge and modifying the outmost layer of the particle. Hence, surface properties are one

of the most important factors that manage NPs mobility and stability as colloidal suspension or

their aggregation and deposition in aquatic environment (Navarro, 2008). Thus, understanding

the stability of colloidal suspensions of NPs may be of great importance in describing and

predicting the environmental fate of NPs in water (Brar et al., 2010; Velzeboer, 2008).

Nanoparticles in natural water systems, especially in waters with higher ionic strength, may have

greater stability when compared to those in natural organic matter (NOM) free waters (Batley

and McLaughlin, 2010). However, in waters with high suspended sediment load, association of

NPs with suspended sediment is likely provide a removal mechanism for NPs that could increase

transport to and accumulation in bottom sediment (Batley and McLaughlin, 2010). In such cases

sediment should be considered as main sink and benthic organisms are key receptors for NPs

released into the aquatic environment since significant sedimentation of NPs aggregates can be

expected (Baun et al., 2008; Christian et al., 2008). In addition, natural fibrillar colloids are

likely to increase aggregation due to different binding characteristics as compared to the charge

stabilisation mechanism of humic substances (Buffle et al., 1998). Hence, the interaction of

ENPs with aquatic colloids may strongly influence the environmental fate of ENPs in surface

water (Lead and Wilkinson, 2007).

The word colloid generally refers to suspended particles less than 10 µm in size (Lead and

Wilkinson, 2007). It includes abiotic colloids such as clay, metal oxides, and humic substances,

and bio-colloids like viruses, bacteria, and protozoa. As a sub-group of colloids, NPs might have

similar size-related properties (Christian et al., 2008). Figure 2.2 shows size domains and typical

representatives of natural colloids and NPs (Christian et al., 2008).

Literature review

8

Figure 2.2 Size domain and typical representatives of natural colloids and nanoparticles;

filtration at 0.45 µm is given as the operationally defined cut-off (Christian et al., 2008).

2.4. Processes affecting the environmental fate of engineered nanoparticles (ENPs)

2.4.1. Aggregation and agglomeration

Aggregation is the association of colloidal particles to form larger clusters that cannot be easily

disrupted (Chen et al., 2010; Gosen et al., 2010). Nanoparticles can aggregate in water through

Van der Waals interactions, chemical bonding, hydrophobic effects, and magnetic attraction.

Coating nanoparticles may decrease aggregation by charge stabilization or steric stabilization.

Aggregation may occur as homoaggregation (particles of the same type aggregating together) or

as heteroaggregation (particle attracting to other particle type) (Handy et al., 2008). Transport

and attachment are the two stages in aggregation; therefore, ENPs and other colloidal particles

must be transported towards each other before aggregation occurs.

Literature review

9

High diffusion coefficients lead to many collisions, and frequent contact between particles

promotes aggregation. The particle concentration can have an effect on the size of aggregates

formed and the speed of aggregation. As the concentration increases, aggregation is more rapid

and aggregates may become large enough to settle out via gravity. So the rate of ENPs

aggregation has an influence on their rate of sedimentation and hence their removal from the

aqueous phase (Chen et al., 2010). The likelihood of permanent attachment is controlled by short

range inter-particle forces of interaction, which in turn depends on solution chemistry, surface

chemistry and composition (Chen, et al., 2010).

Engineered nanoparticles can also agglomerate. Agglomeration is the adhesion of particles to

each other by weaker forces leading to larger size (Gosen et al., 2010). Whereas aggregation is

rather irreversible and implies strong attractive forces, agglomeration is not as strong and more

readily reversible, i.e., agglomerates are easier to break apart into smaller agglomerates or

individual particles. Bare particles often aggregate strongly, whereas surface-coated particles

agglomerate eventually, but can be broken up readily.

Both aggregation and agglomeration of ENPs are processes resulting in a reduction of surface

free energy by increasing their size and decreasing their surface area. Especially aggregated and

agglomerated ENPs can possibly be eliminated through sedimentation, making them less mobile

and leading to interaction with filter feeders and bottom animals. Agglomeration and aggregation

are thus important processes in understanding the fate of nanoparticle in the environment

(Sharma, 2009).

Once released in the environment, ENPs will very likely exist as agglomerated aggregates (Zhou

et al., 2012). Effects of agglomeration and aggregation on the stability and mobility of

nanoparticles have been reported. Zhou et al. (2012) reported that environmental stimuli such as

sunlight and temperature variation can cause either agglomeration or disagglomeration of

(agglomerated) aggregates of metal oxide nanoparticles. Aquatic environments with high

concentrations of calcium or magnesium favour aggregation and deposition of NPs (Fang et al.,

2009).

Literature review

10

2.4.2. Deposition

The process in which ENPs collide and stick over time to an immobile solid surface, such as

sediments, sand, and rocks is referred to as deposition (Hansen et al., 2011). It can take place in

different natural environments, like surface water and groundwater, and is particularly important

in systems where solid surfaces for attachment are readily available. Thus, ENPs deposition is

expected to play a crucial role in influencing the transport of ENPs in aquatic systems.

Aggregation deposition starts when ENPs are transported to a solid surface. Transport

mechanisms for deposition are therefore analogous to aggregation and include Brownian

diffusion, interception, and sedimentation. Both hydrodynamic effects and colloidal interactions

are important factors that influence the adherence of ENPs to solid surfaces (Elimelech et al.,

1995).

2.4.3. Solubility and dissolution

Assessing solubility of NPs is one of the approaches to model their effect, transport, and fate in

the environment (Mackay et al., 2006). The smallest size NPs are energetically unfavourable and

subject to preferential dissolution, and have a higher equilibrium solubility than larger size

particles (Batley and McLaughlin, 2010). Dissolution is a dynamic process that takes place on

the solid-liquid phase boundaries in two steps: a reaction at the solid-liquid interface i.e.

interfacial transport, and transfer of the dissolved matter away from the reaction site

(Dokoumetzidis et al., 2008).

Both particle dissolution kinetics and solubility are size dependent. When NPs are compared

with macro-particles of the same material, NPs dissolve more quickly (Batley and McLaughlin,

2010). A particular concern for metal-based NPs in relation to small size and large surface area is

the dissolution of soluble metal ions from the surface of the particle. Even with solubility of a

few percent, a 1 mg/L solution of metal oxide NPs might generate l µg/L concentrations of metal

ions in solution. Hence, there are concerns that some NPs will act as delivery vehicles for free

metal ions (Handy et al., 2008).

Literature review

11

Sometimes the solubility can exceed saturation conditions and leads to growth and precipitation

of particles by the Ostwald ripening phenomenon, and the overall process is one of

destabilization of NPs in solution. Dissolution and redeposition of particles raise questions about

the overall stability of NPs in the aquatic environment, and also emphasize the need for

measurement of particle size as well as solubility to assess the fate of NPs (Batley and

McLaughlin, 2010).

Furthermore, NPs persistence is governed by its dissolution. In turn, surface area determines the

kinetics of dissolution of soluble material. On the other hand material solubility as well as

metallic ions concentration gradient between particle surface and bulk solution influences the

driving force of dissolution within a given environment (Batley and McLaughlin, 2010; Borm et

al., 2006).

The dissolution of ENPs is extremely important in terms of stability of ENPs as well as the

environmental and human health impact that ENPs may have if released into the environment

(Colvin, 2003). Metallic ENPs dissolution in aqueous environment can give rise to toxic effects

(Zhang et al., 2009), and variation in toxicity may be related to the dissolution properties of

ENPs (Tso et al., 2010).

2.5. Nanoparticles and Contaminant Transport

Contaminants in natural aquatic systems are mostly bound to particle surfaces or form complexes

with humic or other substances (Christian et al., 2008). They can be adsorbed, absorbed, co-

precipitated or trapped upon aggregation of NPs and thus, NPs play an important role in the

solid/liquid partitioning of contaminants. Contaminant sorption onto NPs depends on NPs’

composition, structure, size, and solution conditions like pH and ionic strength. For instance,

pure TiO2 NPs have a higher sorption capacity than impure ones (Christian et al., 2008).

Dispersion of NPs in the environment indicates a higher mobility and a greater potential

exposure to associated contaminants because well dispersed NPs will be transported over longer

distances, and are thus potentially involved in particle-facilitated contaminant transport (Zhuang

et al., 2003).

Literature review

12

2.6. Reactivity of NPs

If a significant change in the atomic structure, electronic, magnetic, and optical properties of the

material is observed, the chemical reactivity of the particle will also be significantly affected.

The factors that contribute to size dependent change in chemical reactivity and properties of a

material can be explained through the following interrelated mechanisms:

(1) Size reduction: the proportion of atoms at the surface or near surface regions increases

considerably when NPs size decreases. This causes a more reactive surface.

(2) Change in surface free energy: the increasing reactive surface leads to a change in surface

free energy with respect to particle size thus influencing the chemical reactivity.

(3) Atomic structure variation: when the size decreases, defects on and near the surface in the

form of change in vacancies, bond length, and bond angle will occur.

(4) Change in electronic structure: as the size gets smaller and smaller the electronic structure

resembles discrete energy states of small molecules (Wigginton et al., 2007).

Changes in size dependent reactivity of material is shown in Figure 2.3 as described by

Wigginton et al., (2007)

Figure 2.3 Generalized trend for size-dependent reactivity change of a material upon transition

from macroscopic (bulk-like) particles to atomic clusters (Wigginton et al., 2007).

Literature review

13

The physicochemical properties of NPs may be different depending on their source and this may

affect NPs’ interactions with organisms. NPs from unintentional sources are mostly poly-

disperse or heterogeneous and irregularly shaped whereas ENPs are mono-disperse or

homogeneous and regular in shape (Sioutas et al., 2005).

2.7. Ecotoxicity of metallic nanoparticles

The toxic effects of metallic nanoparticles are probably due to:

§ Direct release of metals in solutions i.e. dissolution processes;

§ Catalytic properties of metallic nanoparticles;

§ Redox evolution of the surface which can oxidize proteins, generate reactive oxygen

species (ROS) and induce oxidative stress (Auffan et al., 2009; Brunner et al., 2006).

A relationship between the reduction/dissolution of metallic oxide NPs and their ability to

generate oxidative stress has been explained (Limbach et al., 2007). The reduction reaction is

favoured by the corresponding oxidized and reduced molecules present in the biological media.

This can lead to the dissolution of metallic NPs and the release of ions accompanied by the

generation of ROS and oxidation of proteins. These dissolution processes are of great interest in

applications based on the biocide properties of NPs. For instance, antibacterial and antifungal

activities of Ag0 NPs depend on the release of Ag+ ions in solution (Hussain et al., 2005) and the

binding of the surface atoms to electron donor groups containing sulphur, oxygen or nitrogen

(Kumar and Munstedt, 2005; Morones et al., 2005).

Furthermore, the driving force in developing NPs based catalysts is their large surface to volume

ratio and the special binding sites. For example, TiO2 nanoparticles are one of the most

commonly used photocatalysts (Zaleska, 2008). They are efficient to initiate light induced redox

activity with molecules adsorbed on their surfaces. It was found that TiO2 exhibits size

dependent photocatalytic activity and have size dependent biological effects because TiO2 are

highly sensitive to phase transformation as the size decreases (Suttiponparnit et al., 2011). These

size dependent phase transformations are involved not only in the photocatalytic activity of TiO2

particles (Jang et al., 2001) but also in the toxicity of TiO2 towards cellular organisms (Sato and

Taya, 2006).

Literature review

14

Fe0 and Fe3O4 based NPs can become oxidized in biological media. These NPs are used

particularly in the biomedical and environmental fields due to their high sensitivity towards

oxidation and efficiency to degrade organic pollutants (Zhang, 2003). However, recent

nanotoxicological studies report that this oxidation can be responsible for their toxicity towards

environmental bacteria (Auffan et al., 2008b).

Generally, particle size and specific surface area can be key determinants of NPs toxicity. In

addition, other properties such as shape, morphology, crystal structure, composition, purity,

surface chemistry, and particle reactivity may also play a significant role (Tiede et al., 2009).

NPs with more edges have been shown to have higher toxicity to exposed murine fibroblasts and

macrophages (Yamamoto et al., 2004).

Crystal structure may be the basic reason for particle-shape dependent toxicity. For example, the

crystal structure of titanium dioxide NPs controls its ability to induce pulmonary inflammation

and fibrosis in mice (Warheit et al., 2008). Furthermore, particle composition and surface

chemistry may be important in determining biological effects, and NPs’ interaction with

contaminants like trace metals may enhance their toxicity (Gaiser et al., 2012). Because of such a

relationship between NPs properties and toxicity, the quantification of the aforementioned

properties has been recommended as a minimum requirement prior to any toxicological studies

(Baer et al., 2007).

Nanoparticles can enter cells by diffusing through cell membranes (Verma et al., 2008) as well

as through endocytosis (Iversen et al., 2011) and adhesion (Geiser et al., 2005). Some NPs are

intentionally designed to interact with proteins, nucleic acids, or cell membranes for labelling or

drug delivery purposes (Klaine et al., 2008). Additionally, bacteria can be used to deliver NPs

(Demir et al., 2007). However, unintentional interactions are more relevant to environmental

impacts because they are not controlled and they could adversely impact organisms. The

probable toxicity due to NPs uptake and accumulation includes formation of reactive oxygen

species, disruption of membranes or membrane potential, oxidation of proteins, interruption of

energy transduction, genotoxicity, and release of toxic constituents (Klaine et al., 2008).

Literature review

15

For instance, silver NPs adhered to the surface of the cell alter the membrane properties, hence

affecting the permeability and the respiration of the cell. In addition, they can penetrate inside

bacteria and cause DNA damage as well as release toxic Ag+ ions (Klaine et al., 2008).

2.8. Silver nanoparticles (Ag NPs) and silver ions

The use of silver nanoparticles containing consumer products has become common. Major

applications include coatings for solar energy absorption, catalysis in chemical reactions (Choi et

al., 2008), surface-enhanced Raman scattering for imaging, and antimicrobial sterilization (Pal et

al., 2007).

Products containing silver nanoparticles to induce antimicrobial effects are increasingly used in

Europe, North America, and Asia. Because of the high use of consumer products containing Ag

NPs, it is likely these particles enter sewage pipes lines and wastewater treatment plants and

eventually get into surface water (Limbach et al., 2008).

In the European Union for example, 15% of the total silver released into water in 2010 was

predicted to originate from biocidal plastics and textiles (Blaser et al., 2008). However, the

fraction of effectively treated wastewater determines the amount of silver reaching surface

waters, because the majority of silver in wastewater is incorporated into sewage sludge (Blaser et

al., 2008). Due to its strong complexation with different ligands like chloride, sulfide, thio-

sulfate, and dissolved organic carbon, the aqueous concentration of silver ions (Ag+) is low in

natural environments; thus, Ag toxicity to organisms is generally not observed. However, Ag

NPs could be more reactive and toxic than bulk parent material because toxicity is assumed to be

size and shape dependent (Pal et al., 2007). Thus, the potential toxic impact on ecosystems and

microorganisms are major environmental concerns in regard to the release of Ag NPs into the

environment (Blaser et al., 2008; Choi et al., 2008).

2.9. Cerium dioxide nanoparticles (CeO2 NPs)

Large amounts of cerium dioxide nanoparticles (CeO2 NPs) are expected to enter the

environment because of the increasing use of CeO2 in automotive industry as diesel fuel additive

and constituent of catalytic converters (Li et al., 2011).

Literature review

16

Its oxygen storage capacity, the low redox potential between Ce3+ and Ce4+, and UV absorbing

potential are extensively being exploited (Nakagawa et al., 2007). Even though CeO2

nanoparticles are on the OECD list of priority nanomaterials for immediate testing (Van Hoecke

et al., 2003), their environmental fate and potential impacts still remain unclear. CeO2 NPs are

widely applied in polishing materials, automobile exhaust catalysts, as fuel cell materials, and

additives in glass and ceramic application (Van Hoecke et al., 2003). However, CeO2 NPs have

already shown to induce significant chronic toxicity in algae, and to accumulate in zebrafish liver

tissue. It was also reported that human lung fibroblast cells can rapidly absorb CeO2 NPs even at

low concentrations (Limbach et al., 2005).

In addition, CeO2 NPs can produce significant oxidative stress in human lung cells and cause cell

membrane damage. Still, there is no sufficient knowledge on the transport, stability, mobility,

and deposition of CeO2 NPs in the environment, and thus their potential exposure risk, yet.

Hence, understanding the environmental fate and behavior of CeO2 NPs is important to close this

knowledge gap (Li et al., 2011).

2.10. Other common metal(lic) ENPs

2.10.1. Titanium dioxide nanoparticles (TiO2 NPs)

Currently TiO2 NPs are used as catalysts, in antimicrobials, antifungals, antibiotics, ultraviolet

blockers, as antiscratch additives, colour additives in food, drugs, cosmetics, in contact lenses,

and as scavengers of inorganic and organic contaminants in water treatment plants and in the

remediation of polluted environments. They are usually found in soaps, plastics, sunscreens,

coatings, nanofibers and nanowires, textiles, bandages, and alloys (French et al., 2009). This

wide range of applications induces public concern about the possible impacts TiO2 NPs could

have on the environment, for instance as a result of accidental spills during transport and

manufacturing or their presence in waste, sewage, and runoff (French et al., 2009). TiO2 NPs

have been reported in municipal wastewater treatment plant effluents at concentrations of 10 to

100 µg/L (Chen et al., 2011). These particles can eventually reach surface waters and can

possibly accumulate in the environment. So also for TiO2 NPs it is essential to understand their

fate and behaviour in order to determine their bioavailability and toxicity (Chen et al., 2011).

Literature review

17

Furthermore, it was reported that TiO2 could produce reactive oxygen species and cause

oxidative stress in bacteria, crustaceans and various mammal cell types (French et al., 2009).

Surface charge and solution pH mainly govern the stability of TiO2 NPs in aqueous solution.

Over 80% of suspended TiO2 NPs were found to be mobile in micro-channels in the pH range of

1-12, excluding the pH close to the zero point of charge for TiO2 (Guzman et al., 2006).

2.10.2. Iron nanoparticles (Fe NPs)

Fe NPs are synthesized from Fe (II) and Fe (III) using borohydride as reductant (Zhang and

Elliot, 2006). Naturally, iron exists in the environment as iron (II) and iron (III) oxides (Li et al.,

2006). Iron nanoparticles can be used in power transformer cores and magnetic storage media as

well as for catalysis (Guo et al., 2001). Magnetic iron oxide nanoparticles with appropriate

surface modification are also used in biomedical, for example, magnetic resonance contrast

media and therapeutic agents in cancer treatment (Akbarzadeh et al., 2012). Because of the large

surface area of nanoparticles and higher number of reactive sites, nano zero-valent iron (nZVI) is

generally preferred for environmental remediation over microsized particles (Tratnyek and

Johnson, 2007). Moreover, zero-valent iron can also be modified according to the contaminants

to be removed during remediation. For instance, ZVI could be modified to include catalysts like

palladium, coatings such as polyelectrolytes or triblock polymers (Saleh et al., 2007).

However, there is insufficient data available on the potential accumulation in organisms and on

toxicological effects in the environment (Kreyling et al., 2006). Their transport, dispersion, and

fate in the environment is also not sufficiently clear yet. It is known that coatings and other

modifications could maximize subsurface mobility of nZVI (Phenrat et al., 2008). Although

increased mobility does contribute to remediation efficiency, it also leads to the possibility of

NPs migrating beyond the contaminated plume area, thereby leaching into drinking water

aquifers, and wells, and discharge in to surface water might occur (Rajan et al., 2011).

Furthermore, due to their small sizes the nanoparticles have the potential to migrate or

accumulate in areas where larger ones would not (Rajan et al., 2011). Therefore, there is a need

to understand the transport, aquatic and biochemistry, and ultimate fate of these synthetic iron

nanoparticles in the environment (Zhang and Elliot, 2006).

Literature review

18

2.10.3. Gold nanoparticles (Au NPs)

The application of gold colloids goes back a very long way in time where they were used to

create a dark red coloration in glass (Capek, 2004). However, production of Au NPs boomed in

the 20th century due to their wide range of uses. Au NPs possess unique electronic and optical

properties (Daniel et al., 2004). Currently they are used in dyes, inks, films, as catalysts, in drug

delivery, and imaging (Klaine et al., 2008; Renault et al., 2008; Simon-Deckers et al., 2008).

Thus, there is a potential chance of Au NPs being released into the environment throughout its

lifecycle. When compared to their bulk form Au NPs act differently. Many factors such as size,

shape, shell or surface chemistry contribute to Au NPs specific behaviour (Simon-Deckers et al.,

2008). Although knowledge on Au NPs’ toxicity in aquatic environments is not sufficiently high,

there are some reports indicating the probability of its toxicity (Renault et al., 2008; Moor et al.,

2006).

2.10.4. Tin dioxide nanoparticles (SnO2 NPs)

Metal oxide NPs have shown a great step to functionalize materials with high surface to volume

ratio, leading to enormous effective applications such as gas sensors (Phadungdhitidhada et al.,

2011). Semiconducting metal oxide are timely options for gas sensing application because of low

cost, easy fabrication, high compatibility with different parts and processes, and high sensitivity

towards target gas (Arafat et al., 2012). Tin dioxide (SnO2) is one of the rapidly growing NPs in

this group of metal oxides. It is an n-type semiconductor with a large band gap of 3.6 eV

(Farrukh et al., 2010). It has wide range of applications such as transistors, gas sensors, energy

storage like lithium batteries, transparent conducting electrode, and solar cells ( Pirmoradi et al.,

2011). Tin was detected in the gills, gut, and spleen of guppy fish kept in water containing SnO2

NPs (Krysanov et al., 2009). However, hydrated tin dioxide NPs did not cause any acute toxicity

or genotoxicity in short term tests (Krysanov et al., 2009). Nevertheless, the increased tin

contents of the gills and gut suggest that tin nanoparticles may penetrate from the environment

into the body through these organs. In any ways of penetration into the body, tin nanoparticles

occurred in the blood, which is evidenced by their considerable accumulation in the spleen. This

in turn shows that the mobility of SnO2 NPs in the environment and thus their behaviour needs to

be studied.

Literature review

19

2.11. Sediment

Sediment is particulate material such as sand, silt, clay or organic matter that has been deposited

on the bottom of a water body and is susceptible to being transported by water (Stronkhorst et

al., 2004). It is a main source of nutrients for organisms and provides a habitat for benthic

animals. The chemical composition of stream sediments depends on morphology, structural

setting, and lithology of the catchment, climate, hydrological features as well as the density and

type of vegetation cover (Hwang et al., 2011). In addition, anthropogenic activities can influence

bed load and suspended sediment dynamics and the environmental quality of the sediment

systems (Hwang et al., 2011).

Hazardous contaminants such as heavy metals accumulate within the sediments of lakes, rivers

and marine areas. In that way, sediments can act as sinks and/ or carriers for pollutants that can

either be transported away from their source or stored in the solid fraction of bed sediments. If

sediments store contaminants they can become a contaminant source if changes occur in the

environmental conditions within the sedimentary column or in the river course as well as if the

solids are removed and re-suspended (Dinelli et al., 2005). Under anoxic conditions within the

sediments, contaminants are often strongly bound to the solid phase, but once exposed to an oxic

environment the contaminants may be released, and can become bioavailable and or toxic

(Simpson et al., 1998).

Materials and methods

20

3. MATERIALS AND METHODS

3.1. Introduction

Experiments to study partitioning of CeO2 and Ag ENPs were set up in batch mode. The

experiments were conducted using sediment samples from the River Scheldt (Antwerp region).

The sediment samples were analyzed for pH-H2O, pH-KCl, pH-CaCl2, % dry matter (DM), %

organic matter (OM), electrical conductivity (EC), total nitrogen (TN), total phosphorous (TP),

Na, Ca, Mg, K and trace elements, cation exchange capacity (CEC), Cl-, dissolved organic

carbon (DOC), and texture, before being utilized in the batch experiments.

3.2. Sediment sampling and characterization

3.2.1. Sediment sampling and preparation

Sediment samples (ca. 0 – 30 cm) were collected through grab sampling at four different

locations along the River Scheldt: Doel (1), Linkeroever (2), Bornem (3), and Mariekerke (4), as

shown in Figure 3.1.

Figure 3.1 Sediment sampling sites; Doel (1), Linkeroever (2), Bornem (3), and Mariekerke (4)

4

3

2

1


21

After collection, the sediment samples were transferred to the laboratory in sealed plastic bags.

Once arrived in the lab the samples were first air-dried in a greenhouse at 25 ºC, followed by

further drying in an oven at 65 ºC. Afterwards, the different sediment samples were crushed or

grinded and finally passed through a 2 mm mesh sieve.

3.2.2. Sediment characterization

In most cases, the Manual for the Soil Chemistry and Fertility Laboratory (Van Ranst et al.,

1999) was used as a reference guide for the determination of the physicochemical characteristics

of the different sediment samples. All measurements were performed in triplicate unless stated

differently.

3.2.2.1. pH determination

Sediment pH values were obtained by measurement of the samples in three different media:

H2O, KCl and CaCl2.

§ pH-H2O

50 mL of distilled water was added to a glass beaker containing 10.00 grams of sediment. The

suspension was stirred using a glass rod and then left to equilibrate for 16 hours. After

equilibration the pH was measured using a pH electrode (model 520A pH meter, Orion Research

Inc., Boston, MA, USA).

§ pH-KCl

25 mL of 1M KCl was added to a glass beaker containing 10.00 grams of sediment. The

suspension was stirred using a glass rod and then left to equilibrate for 10 minutes. After



§ pH-CaCl2

25 mL of 0.01M CaCl2 was added to a glass beaker containing 10.00 grams of sediment. The

suspension was stirred using a glass rod and then left to equilibrate for 30 minutes. After




22

3.2.2.2. Dry matter (DM)

A known amount of sediment was oven dried for 24 hours at 105 0C and dry matter content was

determined as a weight difference and expressed as percentage using the following formula:

)1.3.(100*)][1(0

10 eqmmm

DM−

−=

Where: DM is the dry matter content (%), m0 is the mass of sediment before oven drying (g),

m1 is the mass after drying in the oven (g).

3.2.2.3. Organic matter (OM)

Organic matter content was determined as loss on ignition (LOI) after ashing of 3.000 g oven

dried sediment sample in a muffle furnace at 550 ºC for 2 hours. Equation 3.2 is used to calculate

the percentage of organic matter content.

)2.3.(100*)(0

10 eqmmm

OM−

=

Where: OM is the total organic matter content (%), m0 is the mass of sediment before ashing (g),

m1 is the mass after ashing (g).

3.2.2.4. Electrical conductivity (EC)

The electrical conductivity was determined by adding 50 mL of distilled water to 10 grams of

sediment in an Erlenmeyer flask and shaking this mixture for 30 minutes on a shaking plate.

Subsequently, the suspension was filtered using filter paper and the filtrate was analyzed directly

using a conductivity meter (LF537, WTW, Weilheim, Germany).

3.2.2.5. Cation exchange capacity (CEC)

Three gram of sediment was thoroughly mixed with 12.5 g pre-cleaned silica sand, poured into a

glass percolation column already containing 2.5 to 3 g pure silica sand, and finally covered up

with again 2.5 to 3 g pure silica sand. After percolation with 150 mL 1 M ammonium acetate, the

excess NH4+ is washed away with 150 mL ethanol solution. Finally, the NH4

+ ions are desorbed

by percolating 250 mL 1 M KCl.


23

This percolate is collected in 250 mL volumetric flasks. Afterwards, 50 mL of the collected KCl

extract is transferred into a Kjeldahl distillation flask and distilled immediately after addition of a

spoonful of MgO. The ammonia (NH3) formed during distillation is captured in the form of

ammonium (NH4+) in an Erlenmeyer flask already containing 20 mL 2 % boric acid (indicator

solution). This solution is then titrated with 0.01 N HCl by means of Metrohm 645 Multi-

Dosimat (Metrohm, Switzerland) until the color shifts to pink. CEC values are calculated in

meq/100 g using the following formula.

)3.3.()100(

100*

250*

1**)( 0 eq

gg

MmL

VtVVCEC

ss

−=

Where: CEC is cation exchange capacity (meq/100 g), V is volume of HCl added to the sample

(mL), V0 is volume of HCl added to the blank (mL), t is normality of HCl in meq/mL, Vs is the

distilled volume of KCl extract (50 mL), Ms is sample mass.

3.2.2.6. Chlorides (Cl-)

Fifty mL 0.15 M HNO3 was added into an Erlenmeyer flask containing 10 g of sediment and the

mixture was left shaking on a shaking plate for 30 minutes. The suspension was filtered over

filter paper and the filter was then rinsed with 20 mL 0.15 M HNO3. The amount of chlorides

present in the sediment sample is determined via a potentiometric titration of the extract with

silver nitrate (AgNO3) using a Metrohm 718 STAT titrino apparatus (Metrohm, Switzerland),

after standardization of AgNO3 with 0.01 N NaCl. The chloride concentration of the sediment

sample is calculated using the following formula:

)4.3.(**

eqM

MCVCl

s

w=−

Where: V is volume of silver nitrate used during titration, C is normality of silver nitrate, Mw is

atomic weight of chloride, MS is sample mass.

3.2.2.7. Total nitrogen (TN)

About 1.000 g of sediment and 7 mL of a mixture of sulphuric/salicylic acid were added into a

glass digestion flask and allowed to react for 30 minutes. Afterwards, 0.5 g Na2SO3was added

and allowed to react for 15 minutes.


24

Then, consecutively 5 mL H2SO4, 0.2 g selenium reagent mixture catalyst and 4 mL H2O2 were

added. The mixture was digested for 1 hour at 380 ºC and afterwards 30 mL of distilled water

was added to the cooled digested product. Finally, distillation was performed and the distillate

was captured in boric acid which is then titrated with 0.01 M HCl using a 645 Multi-Dosimat

titrating apparatus (Metrohm, Switzerland). The expression to determine the total nitrogen

content was:

)5.3.(**)( 0 eqMtMVV

TN ws

−=

Where: TN is total nitrogen (mg/kg), V is volume of HCl used to titrate the sample, V0 is

volume of HCl used to titrate the blank, t is normality of HCl (meq/mL), Mw is molecular

weight of nitrogen, MS is sample mass.

3.2.2.8. Na, K, Ca, Mg, and trace metal contents

Pseudo-total analysis was performed via inductively coupled plasma optical emission

spectrometry (ICP-OES) after aqua regia digestion of the different sediment samples. Therefore,

1.000 g sediment sample was moistened with 2.5 mL distilled water in an Erlenmeyer flask.

Then, 10 mL aqua regia (7.5 mL HCl and 2.5 mL HNO3) was added and the recipient was

covered up with a watch glass. The sample was allowed to digest at room temperature for 12

hours followed by boiling under reflux on a hot plate at 150 ºC for 2 hours. After the sample was

allowed to cool down sufficiently, it was filtered on an acid resistant filter into a 100 mL

volumetric flask. Finally, 1 % HNO3 was added to the filtrate to reach 100 mL and the samples

were analysed by ICP-OES (Vista-MPX CCD Simultaneous ICP-OES, Varian, California).

3.2.2.9. Total phosphorous (TP)

The determination of TP contents of sediments extracted with the solutions NH4OAC-EDTA and

aqua regia was carried out according to the colorimetric method of Scheel. One mL extract, 5 mL

water, 1 mL Scheel solution I (1 g methanol, 5 g Na2SO3 and 150 g NaHSO3), and 1 mL Scheel

solution II (50 g ammonium molybdate and 140 mL concentrated sulphuric acid) were

successively transferred into test tubes, shaken for perfect homogenization and allowed to react

for 15 minutes.


25

Next, 2 mL Scheel solution III (205 g sodium acetate in 1 L water) was added, and the mixture

was shaken and allowed to react for an additional 15 minutes. Finally, the absorbance at 700 nm

was measured with a 6400 spectrophotometer (Jenway, UK).

3.2.2.10. Total carbonate (CO3 2-)

Determination of CaCO3 percentage in the sediment samples is based on the following reactions:

OHCaSOCOSOHCaCO 242423 ++⇔+

OHSONaSOHNaOH 24242 22 +→+

An excess amount (25 mL) of 0.25 M H2SO4 was added to 1.000 g sediment together with 125

mL distilled water. This mixture is then heated for 1 hour at 90 ºC by means of a water bath.

After cooling down, 0.5 mL of mixed indicator solution (phenolphthalein, methyl red,

bromocresol green) is added and the excess acid was titrated using 0.5 M NaOH. The percentage

CaCO3 is calculated using the following expression:

)6.3.(5**)V (% 03 eqCVCaCO −=

Where: V0 is NaOH volume used to titrate the blank, V is NaOH volume used to titrate the

sample, C is NaOH concentration.

3.2.2.11. Dissolved organic carbon (DOC)

Distilled water as well as KNO3 were used as extracting agents in the determination of soluble

organic constituents in the sediments. Fifty mL of distilled water or 50 mL 2 mM KNO3

extraction liquid was added to a 100 mL conical flask already contain 5 g sediment. The mixture

was put on a shaking plate for 24 hours. Afterwards, the suspension was filtered over either filter

paper or a syringe filter (0.45 µm). Finally, the filtrate was diluted five times with distilled water

and was analysed on a total organic carbon (TOC) analyser (TOC-VCPN, Shimadzu, Kyoto,

Japan).

3.2.2.12. Texture analysis

The Bouyoucos hydrometer method (Goh et al., 2009) was used to determine the texture of

sediment samples. First, 40 g of sediment sample was pre-treated with H2O2 to oxidize and

remove OM present in the samples.


26

Second, sediment dispersions were prepared by quantitatively transferring the pre-treated

samples into 1 L plastic recipients and adding 100 mL dispersing agent (40 g

sodiumhexametaphosphate and 10 g soda, dried overnight at 105 ºC) and 250 mL distilled water.

The dispersions were then shaken thoroughly overnight on a shaking plate. Third, the sediment

dispersions were transferred into 1 L glass sedimentation cylinders and distilled water was added

to reach a total volume of 1 L. The sediment solutions were then allowed to equilibrate

thermally. To maintain a constant temperature throughout the measurements a warm temperature

bath kept stable at 20 ºC was used. Finally, the sedimentation cylinders were capped with rubber

stoppers, shaken thoroughly, and a number of readings were performed over a 24 hours time

interval after carefully lowering the hydrometer into the suspensions. Sample and blank (i.e. 100

mL dispersing agent and 900 mL distilled water) readings were then used to calculate the particle

size distribution of the samples. All measurements were performed at least in duplicate.

3.3. Assessment of impact of centrifugation speed on amount of suspended solids

The impact of centrifugation speed on the amount of suspended matter in the supernatant was

studied. First, to obtain a 1/10 solid to liquid ratio, 30.0 mL of Milli-Q water was added to 3.000

g of sediment in a 50 mL centrifuge tube. These suspensions were then shaken vigorously on a

shaking plate for at least 30 minutes before being subjected to centrifugation for 10 minutes at

different selected rates i.e. 0, 500, 1000, 1500, 2000, and 2500 rpm. The 0 rpm represents 10

minutes of gravitational settling. Afterwards, the centrifuged and settled samples were gently

placed onto a steady horizontal surface to not disturb the sample, and 20.0 mL of the supernatant

was pipetted with an electronic pipette into pre-dried and pre-weighted disposable aluminium

evaporation dishes. The concentration of sediment still present in solution (i.e. in the

supernatant) was determined in function of centrifugation speed, by determining the weight

remaining inside the evaporation dishes after drying in an oven at 105 ºC for at least 6 hours. All

measurements were performed in quadruplicate.


27

3.4. Effect of centrifugation speed on TOC concentration

Sediment suspensions in a 1/10 solid to liquid ratio were prepared and put on a shaking plate for

24 hours. The supernatant collected after gravitational settling, and centrifugation at 500 rpm and

2000 rpm were analysed on a total organic carbon (TOC) analyser (TOC-VCPN, Shimadzu,

Kyoto, Japan).

3.5. Partitioning behaviour of metallic ENPs in sediment suspensions

CeO2 and Ag ENPs suspensions were purchased from Plasmachem GmbH, Berlin. The

concentration of the CeO2 and Ag ENPs stock suspension were 50 g/L and 0.1 g/L, respectively.

These ENPs stock suspensions were further characterized experimentally. Prior to the actual

experiment, screening tests for different experimental options (Figure 3.2) were conducted.

3.5.1. Screening experiments

The screening experiments were subdivided in two parts. The first part was designed to assess

effect of centrifugation speed (Figure 3.2a). The second part aimed to look at the effect of

different filtration options (Figure 3.2b). For these screening experiments, sediment suspensions

were prepared with Milli-Q water in 1:10 solid to liquid ratio. These suspensions were then

spiked either with Ce(NO3)3 or CeO2 ENPs. Blank samples only containing sediment and Milli-

Q water were also included in the tests.

In the centrifugation experiment (Figure 3.2a), suspension was allowed to settle for 10 minutes

without centrifugation (gravitational settling, 0 rpm), or centrifuged at 500 or 2000 rpm for 10

minutes. The concentration of Ce in the supernatant after centrifugation (SN) and in the total

suspensions were determined by ICP-OES after aqua regia digestion of the sample for 2 hours on

a hot plate. Moreover the SN was filtered over a syringe microfilter (0.45 µm) without aqua regia

digestion.

The filtration options used in the filtration experiment (Figure 3.2b) were syringe filtration over a

pore size of 0.45 µm (Chromafil RC from Macherey-Nagel, Germany), use of a paper filter

(Chromafil RC from Macherey-Nagel, Germany) and centrifugal ultrafiltration (Amicon Ultra-4

centrifugal-UF units from Millipore, MA, USA). Therefore, 25 mL of suspension was filtered

over filter paper into 50 mL volumetric flask, rinsed with Mill-Q water and filled up to the mark.


28

Fig. 3.2a

Sediment blank Aliquot Total digestion with AR and spiked suspension ∆t ICP-OES on shaking plate Centrifugation SN digestion with AR Analysis @ 0, 500, or 2000 rpm for 10 minutes SN microfiltration Fig. 3.2b

Sediment blank Total digestion with AR and spiked suspension Aliquot on shaking plate ∆t Centrifugal UF ICP-OES Centrifugal UF Paper filtration Analysis Filtrate digestion with AR Centrifugal UF Microfiltration Filtrate digestion with AR

Figure 3.2 Experimental flow sheet of the screening experiment to assess a) centrifugation speed

effect (SN = supernatant), b) effect of different filtration types (UF is Ultrafiltration, AR is aqua

regia) (t, ∆t is 2 hr and/or 24 hr).

From the filtrate collected after paper filter filtration, 10 mL was taken for aqua regia (AR)

digestion, 4 mL for centrifugal ultrafiltration (UF) and 20 mL for syringe microfiltration (MF).

Ten mL from the filtrate after MF was taken for AR, and 4 mL for centrifugal UF. Part of the

suspensions was also directly subjected to aqua regia digestion for total analysis using (ICP-

OES) and another part to direct centrifugal ultrafiltration.


29

For centrifugal ultrafiltration, 4 mL of solution was transferred to a centrifugal ultrafiltration tube

using an electronic pipette. The filtrate was diluted using an autodiluter and then analyzed with

ICP-OES.

3.5.2. Partitioning experiments

Sediment suspensions were prepared with Milli-Q water in a 1:10 (m/v) ratio in 50 mL

centrifuge tubes (Figure 3.3). These suspensions were then spiked with a known amount of CeO2

and Ag ENPs or their corresponding ions (i.e. Ce3+ and Ag+) to get the final concentration of 10

mg/L and 2 mg/L in centrifuge tube, respectively. In addition, control (solely ions or ENPs in

Milli-Q water) and blank (sediment in Milli-Q water without spiking ions or ENPs) samples

were also included in the setup (Table 3.1).

CeO2 ENPs were diluted from the stock solution whereas Ag ENPs were directly sampled from

the stock solution. Partitioning of CeO2 and Ag was determined based on Cornelis et al. (2010).

Figure 3.3 Centrifuge tubes used for the partitioning experiments

The spiked suspensions were placed on a shaking plate for 0, 2 and 24 hours (0 hours represents

10 minutes after spiking of ENPs or respective ions into the centrifuge tubes). After these

selected equilibration times, the samples were centrifuged at 0, 500, and 2000 rpm (where 0 rpm

represents 10 minutes of gravitational settling). A centrifuge tube was used for every sampling

time and centrifugation speed.


30

Table 3.1 Contents in the centrifuge tubes during the partitioning experiments

Centrifuge tube no. Centrifuge tube contents I II III IV V

Milli-Q H2O + Ions Milli-Q H2O + ENPs Milli-Q H2O + Sediment [1:10 m/v] Milli-Q H2O + Ions + Sediment [1:10 m/v] Milli-Q H2O + ENPs + Sediment [1:10 m/v]

Part of the supernatant (10 mL) was taken using an electronic pipette to be digested on a hot

plate for 2 hours at 150 ºC. Aqua regia (7.5 mL 37 % HCl and 2.5 mL 65 % HNO3) was used for

digesting the Ce spiked samples, while digestion of the Ag spiked samples was performed using

nitric acid and hydrogen peroxide (4 mL 65 % HNO3 and 1 mL 30 % H2O2). After digestion, the

suspensions were filtered over filter paper and the filtrates were collected into 50 mL volumetric

flasks. The collected filtrates were diluted to the mark with Milli-Q water and analysed by ICP-

OES. For the controls, total solution (i.e., directly after shaking) as well as “supernatant” aliquots

(i.e., after gravitational settling or centrifugation) were collected, digested and analysed by ICP-

OES. For blank and spiked sediment suspensions total solution aliquots were collected and

analysed at t0 and t24. The latter was only done for the Ce3+ and CeO2 ENPs spiked samples. The

experimental setup for the partitioning experiment is presented in Figure 3.4. All measurements

were performed in triplicate.

Centrifuge tube ∆t ICP-OES I, II, III, IV, V 10 mL aliquot Digestion with AR analysis on shaking plate ∆t Centrifugation @ ICP-OES 0, 500, or 2000 rpm 10 mL SN aliquot Digestion with AR analysis for 10 minutes

Figure 3.4 Experimental procedure for the partitioning experiments. ∆t represents: t0 = after 10

minutes, t2 = 2 hours, t24 = 24 hours, AR = aqua regia. The scheme indicates the steps towards

the analysis of the total solution and of the supernatant (SN).

Results

31

4. RESULTS

4.1. Sediment characteristics

Scheldt river sediment characteristics are presented in Table 4.1. In general, the pH of the

sediments across the Scheldt river was neutral to slightly alkali with the highest pH at station

Doel (7.89) and the lowest at Mariekerke (7.22) sampling station. The organic matter content

was highest at Mariekerke (9.35%) and lowest at Bornem (3.47%). A higher electrical

conductivity (EC) was observed at Doel (1929 µS/cm) in comparison to the other three sites.

Additionally, the ECs of Linkeroever and Bornem sediments were almost identical. Mariekerke

sediment showed the highest values for TN, P, and DOC, which were 2.89 mg/g, 3.41 mg/g and

879 µg/g, respectively. In addition, the CEC of Mariekerke sediment (24.28 meq/100 g) was the

highest when compared to the other three sampled sediments, and the chloride content was

highest for the Doel sediment.

Table 4.1 Physicochemical characteristics of sample sediments (mean ± SD, n = 3).

Parameter Doel Linkeroever Bornem Mariekerke pH-H2O 7.89 ± 0.06 7.88 ± 0.02 7.56 ± 0.02 7.55 ± 0.03

pH-KCl 7.64 ± 0.01 7.56 ± 0.00 7.43 ± 0.01 7.22 ± 0.00

pH-CaCl2 7.60 ± 0.15 7.64 ± 0.00 7.40 ± 0.04 7.40 ± 0.02

DM (%) 98.15 ± 0.02 97.15 ± 0.12 98.46 ± 0.11 97.18 ± 0.07

OM (%) 4.75 ± 0.10 5.08 ± 0.05 3.47 ± 0.04 9.35 ± 0.13

CaCO3 (%) 14.33 ± 0.98 15.22 ± 0.50 10.10 ± 0.76 10.17 ± 0.50

EC (µS/cm) 1929 ± 26 864 ± 8 861 ±14 930 ± 11

CEC (meq/100 g sed.) 10.72 ± 0.10 12.56 ± 0.19 9.50 ± 0.00 24.28 ± 0.19

P (mg/g) 1.51 ± 0.01 1.54 ± 0.01 1.88 ± 0.05 2.89 ± 0.03

TN (mg/g) 1.20 ± 0.01 1.46 ± 0.09 1.21 ± 0.08 3.41 ± 0.02

DOC

(µg/g)

H2O (PF) 386 ± 15 381 ± 14 320 ± 31 879 ± 3

H2O (MF) 288 ± 7 311 ± 14 241 ± 3 743 ± 7

KNO3 (PF) 326 ± 46 302 ± 16 293 ± 23 764 ± 26

KNO3 (MF) 281 ± 4 312 ± 23 259 ± 1 706 ± 5

Cl- (µg/g) 1632 ± 13 649 ± 1 234 ± 1 138 ± 2

Results

32

The major and trace element contents of the Scheldt River sediments are shown in Table 4.2.

Doel and Linkeroever had similar calcium contents (67 mg/g), where the calcium contents of

Bornem and Mariekerke were also comparable (50 mg/g). Selenium concentrations at all stations

and Co contents, except for Mariekerke sampling station, were below the detection limit. The Cr,

Fe, Pb, and Mn contents of Mariekerke sediment were much higher than those of the sediments

from the other three sampling locations.

Table 4.2 Metal and trace element contents in the sampled sediments determined via ICP-OES

after aqua regia digestion (mean ± SD, n = 3).

Parameter Doel Linkeroever Bornem Mariekerke Na (mg/g) 1.89 ± 0.42 1.17 ± 0.05 0.54 ± 0.00 0.64 ± 0.04

K (mg/g) 3.45 ± 0.33 4.17 ± 0.33 3.18 ± 0.10 5.81 ± 0.02

Ca (mg/g) 66.71 ± 2.48 67.39 ± 1.08 51.72 ± 2.00 54.54 ± 0.85

Mg (mg/g) 5.95 ± 0.26 5.58 ± 0.13 3.94 ± 0.17 6.47 ± 0.13

Al (mg/g) 10.78 ± 0.32 12.10 ± 1.01 8.43 ± 0.59 22.34 ± 0.68

Fe (mg/g) 25.03 ± 0.14 27.34 ± 1.61 19.63 ± 1.06 41.90 ± 1.59

Co (µg/g) < 3.99 < 3.98 < 3.99 5.33 ± 0.73

Pb (µg/g) 41.14 ± 0.57 44.47 ± 0.69 43.57 ± 2.52 90.87 ± 2.62

Cr (µg/g) 20.83 ± 0.34 24.07 ± 0.57 19.21 ± 1.20 41.65 ± 1.05

As (µg/g) 17.12 ± 1.58 18.24 ± 3.33 10.84 ± 1.91 20.70 ± 2.50

Cd (µg/g) 1.79 ± 0.07 2.14 ± 0.14 2.38 ± 0.17 3.70 ± 0.05

Mn (µg/g) 446 ± 2 588 ± 40 449 ± 29 1234 ± 29

Se (µg/g) < 19.97 < 19.90 < 19.97 < 19.85

Ni (µg/g) 15.46 ± 0.11 17.43 ± 0.06 15.08 ± 1.15 31.73 ± 0.51

Cu (µg/g) 24.64 ± 0.66 28.09 ± 0.34 26.24 ± 1.35 61.00 ± 2.22

Zn (µg/g) 194.8 ± 0.6 229.3 ± 3.8 232.6 ± 11.0 471.0 ± 12.3

Ce (µg/g) 27.28 ± 0.76 28.43 ± 0.54 37.91 ± 0.83 25.23 ± 0.07

Ag (µg/g) 0.53 ± 0.06 0.61 ± 0.10 0.54 ± 0.05 1.10 ± 0.06

Results

33

Generally, the concentrations of Na, K, Ca, Mg, Fe, Mn, Al and Zn in the Scheldt River

sediments were higher than those of the other elements. For Fe, Al, Mn and Zn, higher

concentrations were measured in the Mariekerke sediment (41.9 mg/g, 22.34 mg/g, 1.2 mg/g,

and 0.47 mg/g for Fe, Al, Mn and Zn, respectively) compared to the other sediments. The

concentrations of these elements in the sediments collected at Doel, Linkeroever and Bornem

sampling stations were almost similar. The background Ce concentration was highest at Bornem

and almost similar at the other three sampling locations, whereas the Ag concentration was two

times higher at Mariekerke compared to the other sampling stations.

The texture of the sediments is presented in Figure 4.1. A high clay content (27 %) was

measured at Mariekerke, while sediments at the other three sites had comparable lower clay

contents. The sand content at Bornem (61 %) was much higher than the sand content at

Mariekerke (26 %). The silt content was similar at Doel, Linkeroever, and Mariekerke but much

lower in Bornem.

Figure 4. 1 Texture of the sediments sampled at Doel, Linkeroever, Bornem, and Mariekerke

(bars represent mean values, error bars represent SD, n = 3).

0

10

20

30

40

50

60

70

Sand Silt Clay

Perc

enta

ge (

%)

Doel Linkeroever Bornem Mariekerke

Results

34

4.2. Assessment of impact of centrifugation speed

Different centrifugation speeds were tested to assess their impact on the amount of suspended

matter remaining in the suspension. The results are shown in Figure 4.2. As expected, suspended

matter in the supernatant (SN) decreases with increasing centrifugation speed for all of the

sediment samples. The amount of suspended matter after 10 minutes of gravitational settling

(indicated by a centrifugation speed of 0 rpm in Figure 4.2), was a lot lower for Bornem (1.5 %)

than for the other three sediments (2.9, 2.7, 3.6 % for Doel, Linkeroever, and Mariekerke,

respectively). The amounts of suspended matter remaining in the suspension for each

centrifugation speed, except in gravitational settling, were similar in Doel and Mariekerke.

Equally, similar amounts of suspended matter were observed for Linkeroever and Bornem as

indicated in Figure 4.2. However, for all of the tested sediments, less than 1 % of the suspended

matter remained in suspension after centrifugation at a speed of 1500 rpm or more for 10 min.

Figure 4.2 Amount of suspended matter in function of centrifugation speed, after centrifuging

for 10 minutes. The amount is presented relative to the initial amount of material in suspension

prior to centrifugation (C/C0 expressed in %). The rate “0” represents 10 minutes of gravitational

settling; bars represent mean values, error bars represent SD (n = 4).

00.5

11.5

22.5

33.5

44.5

0 500 1000 1500 2000 2500

C/C

0 (%

)

Centrifugation speed (rpm)


Results

35

4.3. Assessment of the effect of centrifugation speed on total organic carbon (TOC) in the

supernatant

Total organic carbon contents in the supernatant (SN) of the sediment suspensions was

determined after centrifugation at different centrifugation speeds (0, 500 and 2000 rpm). The

results are presented in Figure 4.3. For all sediment suspensions, TOC contents in the SN

decreased with increasing centrifugation speed. The gravitationally settled sediment of Doel had

the highest TOC concentration in its supernatant (245.6 mg/L).

Figure 4.3 TOC content in the SN of the different sediment suspensions in function of

centrifugation speed (bars represent mean values, error bars indicate SD, n = 3).

4.4. Screening experiments

Cerium concentration in aqua regia digests of suspensions of blank sediment, control ions,

control CeO2 NPs, and sediment suspension spiked with ion or CeO2 ENPs are presented in

Figure 4.4. It concerns total Ce concentrations in the suspensions befor centrifugation.

Background Ce concentrations released from blank sediments, i.e. sediments not spiked with Ce

ions or ENPs, were quite similar for all sediments. The equilibration time did not influence the

total Ce concentration in suspensions.

.

050

100150200250300

0 500 2000

TOC

(mg/

L)

Centrifugation speed (rpm)


Results

36

Figure 4.4 Cerium concentrations in aqua regia digests of suspensions of blank sediment, control

ions and control CeO2 NPs (without sediment), and sediment suspensions spiked with Ce ions or

CeO2 ENPs (D = Doel, L = Linkeroever, B = Bornem, M = Mariekerke, t2 represent sampling

after 2 hours and t24 represents sampling after 24 hours).

Upon centrifugation at different speeds, the concentration of Ce in the controls (i.e. Ce4+ and

CeO2 ENPs in Milli-Q water without sediment) (Figure 4.5) did not differ from their total

concentration observed after aqua regia digestion (Figure 4.4). This indicates that centrifugation

speed has no effect on Ce ions nor on CeO2 ENPs dissolved or suspended in pure Milli-Q water.

Also no change in concentration over time was seen when comparing Ce concentrations in these

control solutions between t2 and t24. Although background Ce concentrations were observed in

aqua regia digests of non-centrifuged blank sediment suspensions (Figure 4.4), Ce concentrations

in supernatants of all centrifuged blank sediment suspensions were below the detection limit

(0.07 mg/L).

02468

1012141618

Ce

conc

entr

atio

n (m

g/L) t2 t24

Results

37

Figure 4.5 Cerium concentrations in aqua regia digests of supernatants obtained after

gravitational settling or centrifuging Milli-Q water previously spiked with Ce ions or CeO2NPs

at different speeds (t2 represents sampling after 2 hours and t24 represents sampling after 24

hours).

In general, higher Ce concentrations in SN of the spiked sediment suspensions (Figure 4.6) were

observed upon gravitational settling. The maximum Ce concentration in SN after gravitational

settling of the sediments spiked with CeO2 ENPs was measured in Linkeroever (4.5 mg/L) and

the lowest concentration was in Mariekerke (1.9 mg/L) at 2 hours equilibration time. Similarly,

the concentration after centrifugation at 500 rpm was highest in Linkeroever (3.8 mg/L) and

lowest in Mariekerke (1.3 mg/L). Concentrations measured in Doel and Linkeroever sediment

were always similar, as were the concentrations measured for Bornem and Mariekerke sediment.

In the case of sediment spiked with Ce ions the maximum Ce concentration (1.7 mg/L) was

observed in Doel sediment and the minimum (0.2 mg/L) in Mariekerke sediment. The Ce

concentrations in the SN of all sediments spiked with Ce ions and some sediment spiked with

ENPs spiked subjected to centrifugation at 2000 rpm were below the detection limit (Figure 4.6).

0

2

4

6

8

10

12

Ctr Ce ions (t2)

Ctr Ce ions (t24)

Ctr CeO2NPs (t2)

Ctr CeO2NPs (t24)

Ce

conc

entr

atio

n (m

g/L)

0 rpm 500 rpm 2000 rpm

Results

38

Figure 4.6 Cerium concentrations in the supernatant of sediment suspensions previously spiked

with Ce ions or CeO2NPs and subjected to gravitational settling or centrifugation at different

speeds (D = Doel, L = Linkeroever, B = Bornem, M = Mariekerke; t2 represents sampling after 2

hours and t24 represents sampling after 24 hours).

Filtrates obtained through microfiltration (MF) of the supernatant collected after centrifugation at

different centrifugation speeds were analysed directly with ICP-OES (without aqua regia

digestion) and the results are presented in Figure 4.7. Cerium concentrations in blank sediments,

and sediments spiked with Ce ions and CeO2 ENPs were under the detection limit (0.07). Almost

all Ce ions dissolved in Milli-Q water could pass MF membrane, whereas some of ENPs could

not pass (Figure 4.7) as compared with the Ce concentration measured in the aqua regia digests

of supernatant of the control (Figure 4.5).

0123456

Ce

conc

entr

atio

n (m

g/L)


Results

39

Figure 4.7 Cerium concentrations in filtrates obtained through microfiltration (MF) of the

supernatant collected after centrifugation at different centrifugation speeds (direct ICP-OES

analysis without aqua regia digestion).

The effect of different filtration procedures was also checked in the screening experiment, as

previously described in paragraph 3.5.1. The results are shown in Figure 4.8. The control Ce ion

was found to pass the filtrate in each filtration procedure. However, this was not the case for the

control ENPs. Cerium could not be detected anymore when all filtration steps were applied to

Doel sediment suspensions spiked with either CeO2 ENPs or Ce ions (Figure 4.8).

Figure 4.8 Cerium concentration in total suspension, and filtrate after different filtration steps

applied to blank, controls and spiked sediment suspensions (UF is ultrafiltration, PF is filtration

paper, PF-UF is filtration paper followed by ultrafiltration, PF-MF is filtration paper followed by

microfiltration, and PF-MF-UF is paper filtration followed by microfiltration and ultrafiltration;

the different filtration procedures are described in paragraph 3.5.1).

02468

1012

Ctr ion (t2) Ctr ion (t4) Ctr NP (t2) Ctr NP (t24)

Ce

conc

entr

atio

n (m

g/L) 0 rpm 500 rpm 2000 rpm

02468

1012

Blank D Ctr.Ce ions Ctr.CeO2 NPs D Ce ions D CeO2 NPs

Ce

conc

entr

atio

n (m

g/L)

Total solution Total UF PF PF-UF PF-MF PF-MF-UF

Results

40

4.5. Partitioning experiment

The partitioning experiments, described in paragraph 3.5.2, were performed separately in batch

mode using suspensions of the four different sediments in Milli-Q water, to which two types of

ENPs (CeO2 and Ag) or their corresponding ions were spiked. During these partitioning

experiments, Ce and Ag concentrations in the SN collected after different centrifugation speeds

were measured at t0 (which represents 10 min after spiking), t2 and t24 (representing 2 and 24

hours after spiking, respectively) Concentrations in AR digests of the total suspensions (non

centrifuged) were measured at t0 and t24 only for Ce.

The Ce concentrations measured in AR digests of total suspensions (non centrifuged) of blank

and spiked sediment suspensions were quite similar and no change in concentration over time

was seen (Figure 4.9).

Figure 4.9 Cerium concentration in aqua regia digests of blank and spiked sediment suspensions

(D is Doel, D ion refers to Doel sediment suspension spiked with ions, D NP is refers to Doel

sediment suspension spiked with ENPs, L is Linkeroever, L ion refers to Linkeroever sediment

suspension spiked with ions, L NP refers to Linkeroever sediment suspension spiked with ENPs,

B is Bornem, B ion refers to Bornem sediment suspension spiked with ions, B NP refers to

Bornem sediment suspension spiked with ENPs, M is Mariekerke, M ion refers to Mariekerke

sediment suspension spiked with ions, M NP refers to Mariekerke sediment suspension spiked

with ENPs, t2 represents sampling after 2 hours, t24 represents sampling after 24 hours; bars

represent mean value, error bars indicate SD, n = 3).

02468

101214

Ce

conc

entr

atio

n (m

g/L)

t0 t24

Results

41

Similarly, no change in concentration of Ce and Ag was seen between the AR digests of non-

centrifuged and centrifuged controls (Figure 4.10). The concentration was also quite stable over

time.

Fig. 4.10a Fig. 4.10b

Fig.4.10c

0

2

4

6

8

10

12

Ctr ion T Ctr NP T

Ce

conc

entr

atio

n (m

g/L)

t0 t2 t24

1.6

1.7

1.8

1.9

2.0

Ctr ion T Ctr NP TA

g co

ncen

trat

ion

(mg/

L)

t0 t2 t24

0

2

4

6

8

10

12

Ctr ion SN (t0)

Ctr ion SN (t2)

Ctr ion SN (t24)

Ctr NPs SN (t0)

Ctr NPs SN (t2)

Ctr NPs SN (t24)

Ce

conc

entr

atio

n (m

g/L)


Results

42

Fig. 4.10d

Figure 4.10 Cerium and Ag concentrations in aqua regia digests of control ions or ENPs

dissolved or suspended in Milli-Q water: a) Ce control non-centrifuged, b) Ag control non-

centrifuged, c) Ce controls centrifuged (supernatant), d) Ag controls centrifuged (supernatant);

Ctr ion represents control suspensions spiked with ions , Ctr NP represents control suspensions

spiked with ENPs; T represents total suspension before centrifugation, SN represents supernatant

after centrifugation; t0 represent 10 minutes, t2 represent 2 hours, t24 represent 24 hours after

spiking; bars represent mean values, error bars indicate SD, n = 3.

Cerium concentrations in the supernatant after different centrifugation speeds are presented in

Figure 4.11. A higher concentration of Ce (6.75 mg/L) was observed in SN after 2 hours shaking

(t2) upon gravity settling of Linkeroever sediment suspensions spiked with CeO2 ENPs (Figure

4.11b). At t2 and gravitational settling maximum 5 mg/L Ce was measured for Ce ion spiked

Linkeroever sediment suspensions (Figure 4.11a). The Ce concentration in SN at 2000 rpm was

under detection limit (< 0.07 mg/L) for all sediment suspensions spiked with Ce ion (Figure

4.11a). In addition, the concentration of Ce in the SN of Bornem sediment suspensions spiked

with Ce ions or Ce ENPs was also under detection limit at 500 rpm and 2000 rpm, respectively

(Figure 4.11a and Figure 4.11b).

1.4

1.5

1.6

1.7

1.8

1.9

2.0

Ctr ion SN (t0)

Ctr ion SN (t2)

Ctr ion SN (t24)

Ctr NPs SN (t0)

Ctr NPs SN (t2)

Ctr NPs SN (t24)

Ag

conc

entr

atio

n (m

g/L)


Results

43

Fig. 4.11a

Fig. 4.11b

Figure 4.11 Cerium concentrations in the supernatant after different centrifugation speeds: a)

sediment suspensions spiked with Ce ions, b) sediment suspensions spiked with CeO2 ENPs (t0

represents 10 minutes, t2 2 hours, t24 24 hours after spiking; D = Doel, L = Linkeroever, B =

Bornem, M = Mariekerke; bars represent mean values, error bars indicate SD, n = 3).

Silver concentrations in the supernatant after different centrifugation speeds are presented in

Figure 4.12. For Ag, maximum concentrations (0.57 mg/L and 1.20 mg/L) in the SN were

observed at the initial time (t0) upon gravitational settling of Bornem sediment suspensions

previously spiked with Ag ion (Figure 4.12a) and Ag ENPs (Figure 4.12b), respectively.

0

2

4

6

8

Ce

conc

entr

atio

n (m

g/L)


0

2

4

6

8

Ce

conc

entr

atio

n (m

g/L)


Results

44

Fig 4.12a

Fig.4.12b

Figure 4.12 Silver concentrations in the supernatant after different centrifugation speeds: a)

sediment suspensions spiked with Ag ions, b) sediment suspensions spiked with Ag ENPs (t0

represents 10 minutes, t2 2 hours, t24 24 hours after spiking; D = Doel, L = Linkeroever, B =

Bornem, M = Mariekerke; bars represent mean values, error bars indicate SD, n = 3).

Both Ce and Ag concentrations in the SN of blank sediment suspensions (sediment suspensions

in Milli-Q water and not spiked with ion or ENPs) were under detection limit. Generally, there

was a clear variation in Ce and Ag concentration in the SN between ion or ENPs spiked

suspensions of the same sediment type and their corresponding control. In addition,

concentration differences in the SN among the different sediment suspensions previously spiked

with ion or ENPs were seen.

0.00.10.20.30.40.50.60.7

Ag

conc

entr

atio

n (m


0.0

0.5

1.0

1.5

2.0

Ag

conc

entr

atio

n (m


Results

45

There was also difference between the two ENPs used in the partitioning experiment. In the case

of Ce, Doel and Linkeroever showed a quite similar behaviour whereas in case of Ag their

behavior was different, especially in the suspension spiked with Ag ions. Moreover, Bornem

sediment behave differently in Ce and Ag partitioning experiments, both the ion and ENPs

spiked suspension.

Discussion

46

5. DISCUSSION

5.1. Sediment characteristics

The organic matter (OM) content of the sediments measured in this study (3.47 % to 9.35 % )

lies in the range reported by Du Laing et al. (2007) for intertidal Scheldt sediments. Organic

matter in river sediment is mostly derived from primary production within aquatic ecosystems

and also from terrestrial biota by transport of leached and eroded material into the river

(Saravanakumar et al., 2008). The highest OM content (9.35 % ± 0.13 %) at Mariekerke may be

due to the nature of the sediment, high rate of sedimentation, high rate of decomposition of

foliage and other vegetative remains in the sediment (Saravanakumar et al., 2008). Furthermore,

the measured dissolved organic carbon (DOC) content in the sediments was in agreement with

the OM content of the sediment. The highest DOC was observed at Mariekerke and the lowest at

Bornem.

Similar trends were observed in chloride content and EC of the sediment with what was reported

by Du Laing et al. (2007). The chloride content of the sediment near to the sea was > 0.5 g/kg

and further from the mouth, the sediments were not affected anymore by salty water, resulting in

lower chloride contents (Du Laing et al., 2007). The highest chloride concentration was observed

at Doel, which is near to the North Sea and thus influenced by the salty seawater, and the lowest

at Mariekerke which might not have been influenced by the salty water. In relation to distance

from the North Sea, the chloride content of sediments in our study can be plotted as shown in

Figure 5.1.

Discussion

47

Figure 5.1 Chloride content of River Scheldt sediments in relation to distance from the mouth of

the river.

The pH measurements in our investigation agree with those previously reported for Scheldt

sediments (Du Laing et al., 2007), and indicate well buffered sediments. at such high pH (Table

4.1), metal availability is minimal. The metals may directly precipitate as carbonates since the

carbonate content of the sediment is high (10 % to 15%, Table 4.1). This also results in a

decreased metal availability. The highest clay content and OM contribute to Mariekerke

sediment having the highest CEC when compared with the other sampling stations. Equally, the

lowest CEC in sediment of Bornem could also be related with its low clay and high sand content.

This illustrates that clay and OM content highly determine the CEC. Because they may supply

negative charges, any element with positive charges will be attracted and held (Aprile and

Lorandi, 2012). Cations have the ability to be exchanged for other positively charged ions at the

surfaces of clay minerals and organic matter. The most important exchangeable cations are

calcium (Ca2+), magnesium (Mg2+), sodium (Na+), potassium (K+), hydrogen (H+), aluminium

(Al3+) and ammonium (NH4+) (Aprile and Lorandi, 2012). K and Al contents were highest at

Mariekerke, which may be linked to the higher CEC at this sampling site.

Generally, it was observed that sediment in the River Scheldt is contaminated with trace metals

(Table 5.1) because measured contents were above the reference values (Saedeler et al., 2010).

This might be due to the fact that areas along the River Scheldt are densely populated and

industrialized, which has a considerable impact on the river sediment physicochemical

characteristics.

0

500

1000

1500

2000

0 50 100 150

Chl

orid

e co

ncen

trai

ton

(µg/

g)

Distance from the mouth (km)


Discussion

48

Table 5.1 Averages of some of trace metal concentrations observed in sediments of our study

and reference values for comparison (µg/g)

Arsenic Cadmium Chromium Copper Lead Nickel Zinc

Measured 17 1.79 21 25 41 15 195

Reference 11 0.38 17 8 14 11 67

Source of reference values: Saedeler et al. (2010)

Texture of the sediments was categorized using the texture triangle given in Figure 5.2.

Sediments at Doel and Linkeroever were classified as loam, whereas sediments at Bornem and

Mariekerke were classified as sandy loam and clay loam, respectively.

Figure 5.2 Classification of the sediment of Doel (D), Linkeroever (L), Bornem (B), and

Mariekerke using a texture triangle.

5.2. Assessment of impact of centrifugation speed

The impact of centrifugation speed on the amount of suspended matter in the sediment

suspension was studied. Six different centrifugation speeds were selected for the study.

L, D B

M

Discussion

49

The percentage of suspended matter in the supernatant exceeds 2.5 % after gravity settling for 10

minutes for all sediments except Bornem. In the same way, when the samples are subjected to a

centrifugation speed of 500 rpm, the percentage of suspended matter was still greater than 1 %

for all sediments except Bornem. This can be explained by the characteristics of the Bornem

sediment. More than 60 % of the sediment at Bornem was classified as sand (Figure 4.1 and

Figure 5.2), which settles more rapidly upon centrifugation or gravitational settling. The

suspended matter in the supernatant decreases when the centrifugation speed increases. On one

hand, a higher speeds allows also smaller particles to settle. On the other hand, at higher speed

also coarser particles can be formed because of agglomeration during centrifugation, as

previously also suggested by Salim and Cooksey (1981). A decrease in the amount of suspended

matter was observed up to 1500 rpm (Figure 5.3). Beyond 1500 rpm the reduction of suspended

matter was independent of centrifugation speed.

Figure 5.3 Amount of suspended matter staying in the supernatant after centrifugation.

5.3. Screening experiment

Taking the impact of centrifugation on suspended matter into account, three centrifugation

speeds (0, 500, 2000 rpm) were selected for screening experiments with sediments to which

CeO2 ENPs and Ce ions were spiked. The experiments were conducted by sampling suspension

after two different equilibration times (t2 and t24) for each selected centrifugation speed.

0.00.51.01.52.02.53.03.54.0

0 500 1000 1500 2000 2500 3000

C /

C0

(%)

Centrifugation Speed (rpm)


Discussion

50

The preliminary data obtained in screening experiments suggested no difference in Ce

concentrations in the controls (ions or ENPs dissolved or suspended in Milli-Q water) between

non-centrifuged (Figure 4.4) and centrifuged (Figure 4.5) suspension as well as among the

suspensions centrifuged at different speeds. In addition, a change of Ce concentrations over time

was not observed in the controls. This illustrates that Ce ions and ENPs dissolved or suspended

in Milli-Q water are not influenced by either centrifugation or equilibration time. However,

differences in Ce concentrations were observed between the total spiked sediment suspensions

(Figure 4.4) and their respective supernatants collected after centrifugation (Figure 4.6).

Differences seem to be directly related to the amount of suspended matter staying in suspension

during centrifugation. Accordingly, the decreasing amount of Ce in the supernatant with

increasing centrifugation speed may be explained by the fact that spiked ions or ENPs may

associate with suspended sediment particles. These particles are removed from the water column

to an increasing extent when the centrifugation speed increases. Furthermore, the experiment

categorized the sediments in two groups, i.e., Doel and Linkeroever in one group, Bornem and

Mariekerke in the other group, based on the level of Ce measured in the supernatant after spiking

(Figure 4.6). Doel and Linkeroever have similar physicochemical properties, but Mariekerke

exceptionally differs from the other sediments in physicochemical properties. This suggests that

sediment characteristics affect presence of Ce in the supernatant.

For supernatants of control solutions containing Ce ions centrifuged at different speeds,

measured Ce concentrations in filtrates obtained through microfiltration (Figure 4.7) were

similar to those obtained after aqua regia digestion (Figure 4.5). However, for the supernatants

of control solutions containing CeO2 NPs, differences were observed. The measured

concentration was higher upon aqua regia digestion. This change in Ce concentration when using

microfiltration was again confirmed in the screening experiment containing different filtration

steps (Figure 4.8). In the latter experiment, the impact of using paper filtration (PF),

microfiltration (MF) and ultrafiltration (UF) was studied for control solutions (ions or ENPs

dissolved or suspended in Milli-Q water) and spiked Doel sediment. In this experiment, the Ce

concentrations were found to be quite similar when passing the different filtration steps for Ce

ions, but not for control Ce NPs (Figure 4.8).

Discussion

51

This may be explained in two ways. On one hand, some Ce ENPs might be retained on the filters

even though their size was much smaller (6 nm) than the pore size of the filters. Similar types of

losses of metals on MF and UF filtration membranes were previously reported (Guo and

Santschi, 2007). Such loss may be solved by pre-treating the UF and MF filtration membranes

prior to use in the experiment (Guo and Santschi, 2007). In addition, NPs may form aggregates

which are too large to pass through the filters (Navarro et al., 2008). This may lead to an under

estimation of ENPs partitioning and dissolution. Thus, attention should be given to the formation

of aggregates, which needs to be monitored during the partitioning experiment. Generally, it can

be concluded that UF and MF can be used as a pretreatment step to determine the concentration

of soluble Ce ions in solution . Pre-treating the UF and MF membranes may even increase the

suitability of these procedures. However, UF and MF cannot be used as pretreatment step when

assessing dissolved CeO2 NPs concentrations.

5.4. Partitioning experiment

The results obtained in the screening experiments helped to setup the partitioning experiment.

The partitioning experiments were conducted using the same three centrifugation speeds as used

in the screening experiment, but by considering also additional equilibration times, as well as Ag

next to Ce.

Cerium concentrations in total blank sediment suspensions (without Ce ion or Ce ENPs spiked)

and sediment suspension to which Ce ion or Ce ENPs were spiked were similar over time

(Figure 4.9). Moroever, as expected, the Ce and Ag concentration in non-centrifuged and

centrifuged controls (control ion and control NPs without sediment) did not vary much with

equilibration time (Figure 4.10). The absence of the effect of equilibration time was confirmed

through one way ANOVA analysis (p > 0.05). This indicates there was no adsorption on the wall

of the centrifuge tubes.

However, variations in Ce and Ag concentrations as a function of equilibration time were

observed in the supernatant of centrifuged spiked sediment suspensions (spiked either with ion or

ENPs). The concentration differences with respect to centrifugation speed and equilibration time

were statistically significant (p < 0.05) in most cases. This shows that ions and ENPs associate

with the suspended matter and/or aggregate to larger particles, which are affected by

Discussion

52

centrifugation. When the amount of suspended matter in the supernatant decreases due to

increased centrifugation speed (Figure 5.1) the concentration of Ce and Ag especially in

suspensions spiked with ions considerably decreased (Figure 4.11a, Figure 4.12a). This may be

due to adsorption of the ion to the suspended particulate matter, which may cause the ion to be

lost from the liquid phase. Such loss increases when the centrifugation speed increases or when

sufficient time is given for the reaction to occur. At higher centrifugation speed (2000 rpm), for

example, the Ce concentration in the supernatant was found to be under the detection limit (<

0.07 mg/L). A Ce concentration less than 0.07 mg/L was observed even at 500 rpm in the case of

Bornem sediment spiked with Ce ions. This might be directly related with the texture, with 60 %

of the Bornem sediment was sand. This would imply that Ce ions are adhered to sand particles.

However, Ce and Ag concentrations were detected in the supernatant at all centrifugation speeds

and equilibration times when the sediment suspensions were previously spiked with ENPs

(Figure 4.11b and Figure 4.12b). This illustrates that the ENPs are more mobile than the

corresponding ions.

Moreover the stabilizing effect of OM and DOC on ENPs (Fang et al., 2009; . Gao et al., 2009)

might explain the preference of ENPs to stay in the liquid phase to a higher extent compared to

the corresponding ions. After normalized with the total concentration the difference in Ce and

Ag Concentration in the SN of the same sediment suspensions previously spiked either with ions

or ENPs was confirmed by one way ANOVA analysis (p < 0.05).

Differences in Ce and Ag concentrations in the supernatant between the different sediments were

also observed. The differences were statistically significant (p < 0.05, Table 5.2). Comparison of

means indicates statistically significant differences between sampling sites (p < 0.05) at the

initial sampling time for almost all ENPs and ion spiked sediment suspensions (Table 5.2).

However, the difference was in most cases not significant between the sediments from Doel and

those from Linkeroever. This may be attributed to the fact that these two sediments had very

similar characteristics (Table 4.1 and Table 4.2, Figure 5.2). The observed concentration

differences between the different sediment suspensions spiked with ions or ENPs may be linked

to sediment characteristics, as sediment characteristics determine solution properties and solution

properties influence the partitioning (Keller et al., 2010, Baalousha, 2009).

Discussion

53

From Pearson’s correlation analysis (Table 5.3), OM, CEC, DOC, clay, sand, and silt were found

to be negatively correlated with the concentrations in the supernatant, whereas CaCO3, EC, and

Cl- were positively correlated. However, correlations between Ce concentrations in the

supernatant, and OM, DOC, clay and sand are not statistically significant. The latter correlations

were also statistically significant in the case of Ag, but correlations with CaCO3 and Cl- were

not. Aggregation or stabilization of ENPs can occur in the presence of OM (Ottofuelling et al.,

2011). This may be due to combined effects of different sediment properties rather than one

single factor effect (Cornelis et al., 2010b). From the regression analysis, it can be concluded that

CaCO3, EC, TOC and sand can describe the concentration differences in the SN more

dominantly than the other sediment characteristics.

Figure 5.4 presents the concentration of Ce and Ag in the supernatant after gravitational settling

normalised to the total concentration in suspension (before centrifugation) for Ag ENPs and

CeO2 ENPs spiked sediment suspensions. It illustrates that Ce is more mobile in the Doel and

Linkeroever sediments compared to the Mariekerke and Bornem sediments, and compared to Ag.

This may possibly be due to the higher pH, and higher chloride and carbonate contents in these

sediments. Cerium is mobilised to the supernatant first (after 2 hours), but is then immobilised

again (after 24 hours). This was also the case for Ag in the Doel sediment, whereas mobility of

Ag continuously decreased in the other sediments. Silver particles are known to easily release

Ag+ ions. Because Ag+ ions are associated more strongly to particulate material which is more

easily removed from suspension (see above), the dissolution of Ag NPs and coinciding release of

Ag+ ions may have caused this effect.

Discussion

54

a) b)

Figure 5.4 Concentration of Ce and Ag in the supernatant after gravitational settling of ENPs

spiked sediment suspensions, normalised to the total Ce and Ag concentration in suspension

(before centrifugation) (C/C0) for: a) CeO2 ENPs spiked sediment suspension, b) Ag ENPs

spiked sediment suspension; 1 represent t0, 2 represent t2, 3 represent t24.

0.0

0.2

0.4

0.6

0.8

0 1 2 3 4

C/C

0

Time (hours)

Doel LinkeroeverBornem Mariekerke

0.0

0.2

0.4

0.6

0.8

1.0

0 1 2 3 4

C/C

0

Time (hours)

Doel LinkeroeverBornem Mariekerke

Dis

cuss

ion

55

Tab

le 5

.2 M

ean

conc

entr

atio

ns o

f C

e an

d A

g in

sup

erna

tant

s of

spi

ked

sedi

men

t su

spen

sion

s (m

g/L

), sh

owin

g th

e st

atis

tical

sign

ific

ance

of

diff

eren

ces:

a)

Ce

ions

or

Ce

EN

Ps s

pike

d se

dim

ent s

uspe

nsio

ns, b

) A

g io

ns o

r A

g E

NPs

spi

ked

sedi

men

t sus

pens

ion.

Eac

h sm

all a

nd c

apita

l let

ters

den

otes

hom

ogen

ous

subs

et w

ithin

eac

h ra

w a

nd c

olum

n, re

spec

tivel

y.

5.2a

) C

e co

ncen

trat

ion

in s

uper

nata

nt o

f

Sedi

men

ts s

uspe

nsio

ns s

pike

d w

ith C

e io

n

Ce

conc

entr

atio

n in

su

pern

atan

t of

se

dim

ents

susp

ensi

ons

spik

ed w

ith C

e E

NPs

0

rpm

50

0 rp

m

2000

rpm

0

rpm

50

0 rp

m

2000

rpm

t 0

Doe

l 3.

41 ±

0.6

7aE

1.04

± 0

.06bE

5.60

± 0

.06aE

0.

91 ±

0.0

4bE

1.64

± 0

.04cE

Lin

kero

ever

4.

89 ±

0.1

4aF

0.8

± 0.

13bF

5.75

± 0

.05aE

2.

11 ±

0.2

4bF

1.88

± 0

.15bF

Bor

nem

2.

37 ±

0.0

8aG

2.85

± 0

.14aF

0.

78 ±

0.0

2bE

Mar

ieke

rke

2.06

± 0

.18aG

0.

79 ±

0.0

5bF

3.

55 ±

0.1

0aG

2.49

± 0

.19bF

0.

83 ±

0.0

4cG

t 2

Doe

l 4.

49 ±

0.5

0aE

0.41

± 0

.11bE

6.27

± 0

.02aE

2

± 1.

73bE

1.

61 ±

0.0

8bE

Lin

kero

ever

5.

06 ±

0.8

2aE

0.44

± 0

.07bE

6.75

± 0

.10aE

1.

63 ±

0.5

8bE

3.4

± 0.

14cF

Bor

nem

1.

62 ±

0.1

6aF

2.92

± 0

.34aF

1.

13 ±

0.0

5bE

Mar

ieke

rke

1.56

± 0

.14aF

0.

49 ±

0.0

3bE

4.

32 ±

0.1

1aG

1.5

± 0.

35bE

0.

47 ±

0.0

9cG

t 24

Doe

l 3.

33 ±

0.5

8aE

1.17

± 0

.09bE

6.07

± 0

.06aE

2.

51 ±

1.1

3bEF

2.19

± 0

.07bE

Lin

kero

ever

3.

49 ±

0.5

5aE

0.59

± 0

.12bF

6.22

± 0

.08aE

3.

25 ±

0.9

bF

2.72

± 0

.3bF

Bor

nem

1.

08 ±

0.1

6aF

1.54

± 0

.35aF

0.

81 ±

0.0

6bE

Mar

ieke

rke

0.62

± 0

.03aF

2.

67 ±

0.8

5aF

1.29

± 0

.05bE

0.

53 ±

0.0

2bG

Dis

cuss

ion

56

5.2b

)

A

g co

ncen

trat

ion

in s

uper

nata

nt o

f

Sedi

men

ts s

uspe

nsio

ns s

pike

d w

ith A

g io

n

Ag

conc

entr

atio

n in

sup

erna

tant

of

Sedi

men

ts s

uspe

nsio

ns s

pike

d w

ith A

g E

NPs

0

rpm

50

0 rp

m

2000

rpm

0

rpm

50

0 rp

m

2000

rpm

t 0

D

oel

0.32

± 0

.04aE

0.

20 ±

0.0

01bE

0.

19 ±

0.0

1bE

0.44

± 0

.22aE

0.

37 ±

0.0

4aE

0.40

± 0

.09aE

Lin

kero

ever

0.

34 ±

0.0

2aE

0.18

± 0

.01bF

0.

10 ±

0.0

2cF

0.70

± 0

.13aE

F 0.

70 ±

0.0

8aF

1.13

± 0

.43aF

Bor

nem

0.

57 ±

0.0

3aF

0.49

± 0

.01bG

0.

38 ±

0.0

04cG

1.

2 ±

0.02

aF

1.02

± 0

.04bG

0.

70 ±

0.0

3cEF

Mar

ieke

rke

0.38

± 0

.01aE

F 0.

15 ±

0.0

1*10

-1bH

0.

07 ±

0.0

4cF

1.05

± 0

.04aG

0.

64 ±

0.0

1bF

0.35

± 0

.03cE

t 2

Doe

l 0.

30 ±

0.0

2aE

0.15

± 0

.01bE

0.

07 ±

0.0

01cE

0.

89 ±

0.2

4aE

0.28

± 0

.02bE

0.

36 ±

0.0

4bE

Lin

kero

ever

0.

36 ±

0.0

1aF

0.12

± 0

.01bE

F 0.

04 ±

0.0

1cF

0.61

± 0

.04aE

0.

91 ±

0.4

1bF

0.91

± 0

.28cF

Bor

nem

0.

23 ±

0.0

1aG

0.09

± 0

.01bF

0.

03 ±

0.0

1cF

0.89

± 0

.06aE

0.

54 ±

0.0

2bF

0.24

± 0

.01cE

Mar

ieke

rke

0.33

± 0

.01aF

0.

08 ±

0.0

3bF

0.01

± 0

.002

cG

0.79

± 0

.04aE

0.

39 ±

0.0

1bF

0.11

± 0

.01cE

t 24

Doe

l 0.

25 ±

0.0

4*10

-1aE

0.

24 ±

0.0

1aE

0.08

± 0

.01bE

0.

27 ±

0.0

3aE

0.43

± 0

.01bE

0.

39 ±

0.0

9abE

Lin

kero

ever

0.

42 ±

0.0

1aF

0.21

± 0

.02bE

0.

04 ±

0.0

1cF

0.5

± 0.

07aF

0.

76 ±

0.2

2aF

0.62

± 0

.07aF

Bor

nem

0.

31 ±

0.0

4aE

0.09

± 0

.02bF

0.65

± 0

.03aG

0.

26 ±

0.0

3bE

0.08

± 0

.02cG

Mar

ieke

rke

0.18

± 0

.03aG

0.

07 ±

0.0

1bF

0.

34 ±

0.0

3aE

0.17

± 0

.01bE

0.

05 ±

0.0

4*10

-1cG

Dis

cuss

ion

57

Tab

le 5

.3 P

ears

on’s

cor

rela

tion

coef

fici

ent

betw

een

Ce

and

Ag

conc

entr

atio

n in

sup

erna

tant

of

sedi

men

t su

spen

sion

s sp

iked

with

Ag/

Ce

ion

or A

g/C

eO2 E

NPs

and

sed

imen

t cha

ract

eris

tics

A

g E

NPs

A

g io

n C

e E

NPs

C

e io

n O

M

CaC

O3

CE

C

EC

D

OC

C

l C

lay

Sand

Si

lt T

OC

OM

-0

.209

-0

.243

-0

.031

-0

.128

1

Si

g.

0.01

5 0.

006

0.37

5 0.

093

CaC

O3

0.14

9 0.

021

0.50

0 0.

356

-0.3

18

1

Si

g.

0.06

1 0.

416

0.00

0 0.

000

0.00

0

CE

C

-0.1

88

-0.2

33

-0.0

85

-0.1

64

0.99

1 -0

.418

1

Si

g.

0.02

5 0.

008

0.19

0 0.

045

0.00

0 0.

000

EC

-0

.258

-0

.006

0.

198

0.16

6 -0

.180

0.

437

-0.2

91

1

Si

g.

0.00

4 0.

475

0.02

0 0.

043

0.03

1 0.

000

0.00

1

DO

C

-0.2

33

-0.2

31

-0.1

15

-0.1

80

0.98

6 -0

.467

0.

992

-0.2

12

1

Sig.

0.

008

0.00

8 0.

118

0.03

1 0.

000

0.00

0 0.

000

0.01

4

C

l -0

.120

0.

029

0.33

1 0.

266

-0.3

57

0.72

0 -0

.476

0.

932

-0.4

33

1

Si

g.

0.10

8 0.

383

0.00

0 0.

003

0.00

0 0.

000

0.00

0 0.

000

0.00

0

Cla

y -0

.184

-0

.233

-0

.080

-0

.160

0.

992

-0.4

07

1.00

0 -0

.294

0.

991

-0.4

74

1

Sig.

0.

028

0.00

8 0.

207

0.04

9 0.

000

0.00

0 0.

000

0.00

1 0.

000

0.00

0

Sa

nd

0.23

7 0.

251

-0.1

31

0.01

2 -0

.930

-0

.003

-0

.873

-0

.130

-0

.872

0.

001

-0.8

75

1

Si

g.

0.00

7 0.

004

0.08

8 0.

449

0.00

0 0.

488

0.00

0 0.

090

0.00

0 0.

496

0.00

0

Silt

-0.2

51

-0.2

19

0.26

6 0.

103

0.70

9 0.

315

0.60

9 0.

474

0.62

1 0.

394

0.61

2 -0

.918

1

Si

g.

0.00

4 0.

012

0.00

3 0.

144

0.00

0 0.

000

0.00

0 0.

000

0.00

0 0.

000

0.00

0 0.

000

TO

C

0.11

9 0.

366

0.70

8 0.

637

0.46

1 0.

074

0.41

8 0.

170

0.42

3 0.

105

0.41

9 -0

.536

0.

533

1 Si

g.

0.11

1 0.

000

0.00

0 0.

000

0.00

0 0.

223

0.00

0 0.

039

0.00

0 0.

139

0.00

0 0.

000

0.00

0

Conclusion and recommendations

58

6. CONCLUSIONS AND RECOMMENDATIONS

6.1. Conclusions

Batch experiments were set up to examine partitioning behavior of CeO2 and Ag engineered

nanoparticles in suspensions of sediments sampled from four locations along the River Scheldt,

differing in physicochemical characteristics. CeO2 and Ag engineered nanoparticles as well as

their corresponding Ce (III) and Ag (I) ions were spiked into the sediment suspensions. Cerium

and Ag concentrations were analysed in the supernatant after centrifugation of samples taken at

different equilibration times. Prior to this partitioning experiment, experiments were conducted

to assess the impact of centrifugation speed on suspended matter remaining in suspension and to

test the effect of different filtration procedures during sample preparation.

Background concentrations of both Ce and Ag in the supernatant of sediment suspensions were

under the detection limit. The concentration of Ce and Ag in the supernatant previously spiked

with either ENPs or ions significantly depended on equilibration time and centrifugation speed.

Remarkably, CeO2 and Ag ENPs were found to be more mobile, i.e. more present in the

supernatant, than their corresponding Ce (III) and Ag (I) ions. The Ce and Ag concentrations

observed in supernatant differed significantly between different sediment suspensions,

suggesting that sediment properties influence the partitioning behavior of the ENPs. CeO2

nanoparticles were found to be more mobile in the suspensions of two sediments compared to

two other sediments, and compared to Ag nanoparticles. This may possibly be due to the higher

pH, and higher chloride and carbonate contents in these sediments. When spiking CeO2

nanoparticles, Ce is mobilised to the supernatant during the first 2 hours after which it is

immobilised again. This was also the case for Ag in the one sediment, whereas mobility of Ag

continuously decreased in the other sediments. The latter may have been due to a rapid release of

Ag+ ions from the Ag nanoparticles and association of the released Ag+ ions to particulate

material in suspension.

Conclusion and recommendations

59

6.2. Recommendations

Future research is recommended. Trying to find explanations for effects observed in our study,

formation of aggregates as well as dissolution of particles and release of ions from the particles

should be monitored. Including additional sediments and nanoparticles having other properties

may also contribute to this. Moreover, filtration steps may be used to discriminate between ions

and nanoparticles in solution. In this context, pretreatment of the filter membranes should be

investigated as an option to improve recovery of the ions in the filtrate.

References

60

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