Occurrence and environmentalrisk assessment of antifoulingpaint biocides from leisure boats Maria Lagerström

Maria Lagerström

    Occu

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ental risk assessm

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tifoulin

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leisure boats

Doctoral Thesis in Applied Environmental Science at Stockholm University, Sweden 2019

Department of Environmental Science andAnalytical Chemistry

ISBN 978-91-7797-655-4

Occurrence and environmental risk assessment ofantifouling paint biocides from leisure boatsMaria Lagerström

Academic dissertation for the Degree of Doctor of Philosophy in Applied EnvironmentalScience at Stockholm University to be publicly defended on Friday 24 May 2019 at 10.00 inWilliam-Olssonsalen, Geovetenskapens hus, Svante Arrhenius väg 14.

AbstractThe use of biocidal antifouling (AF) paints is the most common way to prevent fouling on leisure boat hulls. The mainaim of this thesis was to investigate the pathways through which AF biocides, past and present, may reach the environmentthrough their use on leisure boats and to improve the risk assessment of biocidal AF paints intended for amateur use. Thework presented focuses mainly on the Baltic Sea, with emphasis on regulation and risk assessment procedures in Sweden.A new method was developed for the quantification of nowadays banned organotin compounds (OTCs) such as tributyltin(TBT) in paint flakes (paper I). OTCs were detected in hull paint scrapings from three countries around the Baltic Sea.Thus, historic layers of organotin paint on leisure boats may constitute as sources of TBT to the marine environment. Totaltin was identified as an indicator for the presence of OTCs on boat hulls, allowing for quicker identification of vessels inneed of remediation. Nowadays, most AF paints tend to contain high amounts of copper (Cu) and zinc (Zn). The use ofAF paints was shown to cause exceedance of guideline values for these two metals in soil, sediment and water in variousinvestigated marinas (papers II and IV). The pollution of boatyard soil was linked to hull maintenance activities carried outover unprotected ground (paper II). AF paints were also found to impact both the concentration and speciation of dissolvedCu and Zn in two Baltic Sea marinas, with increased concentrations as well as an increased proportion of bioavailablespecies as a function of an increased number of moored boats (paper IV). A new method utilizing X-Ray Fluorescence(XRF) was used to derive the release rates of Cu and Zn in the field for five commercially available AF paints for amateuruse (paper III). Salinity and paint properties were found to be important parameters affecting the release. The in situ releaserates were also found to exceed those derived with current standardized release rate methods. Given the high release rates,none of the studied paints should have been approved for the Swedish market. This finding likely explains the exceedanceof guideline values for dissolved Cu and Zn in investigated Baltic Sea marinas (paper IV). In conclusion, there is a needfor caution when authorizing new biocides as the phasing out of banned substances can be a lengthy process due to theircontinued presence in historic paint layers. Additionally, paint-specific release rates determined under conditions reflectingthe intended use of the product should be used for a more realistic environmental risk assessment of AF paints.

Keywords: Antifouling paint, XRF, leisure boat, recreational boating, copper, zinc, tributyltin, organotin compounds,risk assessment, biocide, metals, release rate.

Stockholm 2019http://urn.kb.se/resolve?urn=urn:nbn:se:su:diva-167281

ISBN 978-91-7797-655-4ISBN 978-91-7797-663-9

Department of Environmental Science and Analytical Chemistry

Stockholm University, 106 91 Stockholm

OCCURRENCE AND ENVIRONMENTAL RISK ASSESSMENTOF ANTIFOULING PAINT BIOCIDES FROM LEISURE BOATS 

Maria Lagerström

Occurrence and environmentalrisk assessment of antifoulingpaint biocides from leisureboats 

Maria Lagerström

©Maria Lagerström, Stockholm University 2019 ISBN print 978-91-7797-655-4ISBN PDF 978-91-7797-663-9 Printed in Sweden by Universitetsservice US-AB, Stockholm 2019

Till Andreas

Contents

Abbreviations .................................................................................................. 2

Abstract ........................................................................................................... 3

Sammanfattning .............................................................................................. 4

List of papers ................................................................................................... 5

1 Introduction ............................................................................................ 6

2 Antifouling paints .................................................................................... 8 2.1. Paint types ............................................................................................................ 8 2.2. Antifouling biocides ............................................................................................... 9

3 Aims ..................................................................................................... 12

4 Methods ............................................................................................... 13 4.1. XRF analysis....................................................................................................... 13 4.2. Analytical and field methods ............................................................................... 17 4.3. Release rate methods ........................................................................................ 20

5 Regulation ............................................................................................ 24 5.1. Current regulatory status .................................................................................... 24 5.2. Regulatory challenges ........................................................................................ 26

6 Pollution pathways ............................................................................... 30 6.1. In the harbor ....................................................................................................... 30 6.2. In the boatyard .................................................................................................... 32

7 Risk assessment .................................................................................. 34 7.1. Environmental risk assessment procedure ........................................................ 34 7.2. Towards an improved risk assessment .............................................................. 36

8 Conclusions and future perspectives ................................................... 41

References .................................................................................................... 43

Acknowledgements ....................................................................................... 49

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Abbreviations

AF Antifouling BPR Biocidal Products Regulation CEPE European Council of the Paint, Printing Ink and Artists’

Colours Industry Cu Copper CuCSN Copper thiocyanate Cu2O Cuprous oxide DFT Dry Film Thickness DGT Diffusive Gradients in Thin films EQS Environmental Quality Standard EU European Union ICP-MS Inductively Coupled Plasma Mass Spectrometry ICP-OES Inductively Coupled Plasma Optical Emission Spectroscopy IMO International Maritime Organization MAMPEC Marine Antifoulant Model to Predict Environmental

Concentrations MSFD Marine Strategy Framework Directive OTC Organotin Compound Pb Lead PEC Predicted Environmental Concentration PNEC Predicted No Effect Concentration PSSA Particularly Sensitive Sea Area PSU Practical Salinity Unit Sn Tin TBT Tributyltin TPhT Triphenyltin WFD Water Framework Directive XRF X-Ray Fluorescence Zn Zinc ZnO Zinc oxide

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Abstract

The use of biocidal antifouling (AF) paints is the most common way to prevent fouling on leisure boat hulls. The main aim of this thesis was to investigate the pathways through which AF biocides, past and present, may reach the en-vironment through their use on leisure boats and to improve the risk assess-ment of biocidal AF paints intended for amateur use. The work presented fo-cuses mainly on the Baltic Sea, with emphasis on regulation and risk assess-ment procedures in Sweden. A new method was developed for the quantifica-tion of nowadays banned organotin compounds (OTCs) such as tributyltin (TBT) in paint flakes (paper I). OTCs were detected in hull paint scrapings from three countries around the Baltic Sea. Thus, historic layers of organotin paint on leisure boats may constitute as sources of TBT to the marine environ-ment. Total tin was identified as an indicator for the presence of OTCs on boat hulls, allowing for quicker identification of vessels in need of remediation. Nowadays, most AF paints tend to contain high amounts of copper (Cu) and zinc (Zn). The use of AF paints was shown to cause exceedance of guideline values for these two metals in soil, sediment and water in various investigated marinas (papers II and IV). The pollution of boatyard soil was linked to hull maintenance activities carried out over unprotected ground (paper II). AF paints were also found to impact both the concentration and speciation of dis-solved Cu and Zn in two Baltic Sea marinas, with increased concentrations as well as an increased proportion of bioavailable species as a function of an increased number of moored boats (paper IV). A new method utilizing X-Ray Fluorescence (XRF) was used to derive the release rates of Cu and Zn in the field for five commercially available AF paints for amateur use (paper III). Salinity and paint properties were found to be important parameters af-fecting the release. The in situ release rates were also found to exceed those derived with current standardized release rate methods. Given the high release rates, none of the studied paints should have been approved for the Swedish market. This finding likely explains the exceedance of guideline values for dissolved Cu and Zn in investigated Baltic Sea marinas (paper IV). In con-clusion, there is a need for caution when authorizing new biocides as the phas-ing out of banned substances can be a lengthy process due to their continued presence in historic paint layers. Additionally, paint-specific release rates de-termined under conditions reflecting the intended use of the product should be used for a more realistic environmental risk assessment of AF paints.

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Sammanfattning

Användning av båtbottenfärger som innehåller biocider är det vanligaste sättet att förhindra påväxt på fritidsbåtsskrov. Huvudsyftet med denna avhandling var att undersöka sätten genom vilka gamla och nuvarande biocider kan nå miljön genom deras användning i båtbottenfärger på fritidsbåtar och att för-bättra riskbedömningen av dessa biocidprodukter. Arbetet fokuserar främst på Östersjön, med tonvikt på reglering och riskbedömning i Sverige. En ny metod utvecklades för kvantifiering av numera förbjudna tennorganiska föreningar såsom tributyltennn (TBT) i färgflagor (paper I). Tennorganiska föreningar detekterades i färgavskrap från fritidsbåtar från tre länder runt Östersjön. Gamla färglager av tennorganisk färg på fritidsbåtar kan således utgöra källor av TBT till den marina miljön. Totalt tenn identifierades som en indikator för förekomsten av tennorganiska föreningar på båtskrov, vilket möjliggör en snabbare identifiering av farkoster i behov av åtgärd. Idag innehåller de flesta bottenfärger höga halter koppar (Cu) och zink (Zn). Användningen av båtbot-tenfärger visade sig orsaka överskridande av riktvärden för dessa två metaller i mark, sediment och vatten i olika undersökta hamnar (paper II och IV). Förorening av jorden på båtuppställningsplatser kunde kopplas till skrovun-derhåll som utförts på oskyddad mark (paper II). Båtbottenfärger visade sig även påverka både koncentrationen och specieringen av löst Cu och Zn i två hamnar vid Östersjön, med ökade halter samt en ökad andel biotillgängliga specier som en funktion av ökat antal båtar i hamnen (paper IV). En ny metod som utnyttjar röntgenfluorescens (XRF) användes för att bestämma läckage-hastigheter av Cu och Zn i fält för fem kommersiellt tillgängliga bottenfärger för fritidsbåtar (paper III). Salthalt och färgegenskaper visade sig vara viktiga parametrar som påverkar läckaget. Läckagehastigheterna visade sig dessutom överskrida de som tagits fram med nuvarande standardiserade läckagemeto-der. Givet de höga läckagehastigheterna, borde ingen av de undersökta fär-gerna ha godkänts för den svenska marknaden. Detta resultat förklarar sanno-likt överskridandet av riktvärdena för löst Cu och Zn i de undersökta Öster-sjöhamnarna (paper IV). Sammanfattningsvis finns det behov av försiktighet vid godkännande av nya biocider, eftersom utfasningen av förbjudna ämnen kan bli en utdragen process på grund av deras fortsatta förekomst i gamla färg-skikt. Vidare bör färgspecifika läckagehastigheter bestämda under förhållan-den som återspeglar den avsedda användningen av produkten användas för en mer realistisk riskbedömning av båtbottenfärger.

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List of papers

Paper I Lagerström, M., Strand, J., Eklund, B., & Ytreberg, E. (2017). Total tin and organotin speciation in historic layers of antifouling paint on leisure boat hulls. Environmental pollution, 220, 1333-1341.

Paper II Lagerström, M., Norling, M., & Eklund, B. (2016). Metal contamination at recreational boatyards linked to the use of antifouling paints—investigation of soil and sediment with a field portable XRF. Environmental Science and Pollution Research, 23(10), 10146-10157.

Paper III Lagerström, M., Lindgren, J. F., Holmqvist, A., Dahlström, M., & Ytreberg, E. (2018). In situ release rates of Cu and Zn from commercial antifouling paints at different salinities. Marine Pollution Bulletin, 127, 289-296.

Paper IV Lagerström, M., Ferreira, J., Ytreberg, E. & Wiklund, A.K.E. Flawed risk as-sessment of antifouling paints leads to exceedance of guideline values in Bal-tic Sea marinas. Manuscript.

Author contributions I had the leading role in the writing of all four papers. With regards to data collection, I carried out all the laboratory work (except the GC-PFPD analysis and the acid digestions of paint flakes) and the data analysis for paper I. For paper II, I participated in the planning and execution of all the field work and executed all the laboratory work (except the milling of the samples) and con-sequent data analysis. I participated in the field work, the development of the XRF method and the XRF analyses of samples for paper III. I also performed all the modelling in MAMPEC. For paper IV, I planned all the field work and also participated in the sampling. I also carried out all the chemical analyses (except the DOC analysis) and did the data analysis.

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1 Introduction

Recreational boating is highly dependent on the state of the sea, yet there are numerous ways in which it can negatively impact the marine environment. Noise pollution, emission of sewage water, hydrocarbons and oil, anchor dam-age, littering, propagation of non-indigenous species and emission of toxic substances from antifouling (AF) paints have been identified as the main ways in which boating can affect both marine life and water quality [1,2]. In this thesis, the main focus is on the emission of biocides and other substances of concern to the environment from AF paints used on leisure boats. The work presented here focuses mainly on the Baltic Sea, with emphasis on regulation and risk assessment procedures in Sweden. The Baltic Sea was classified as a Particularly Sensitive Sea Area (PSSA) by the International Maritime Organization (IMO) in 2001 due to its low, but unique biodiversity [3]. This semi-enclosed brackish sea is characterized by a slow water exchange rate (approximately 30 years), making it especially vul-nerable to pollution [4]. With an estimated 1.6 million leisure boats of all sizes frequenting its coastline, AF biocides are of environmental concern to the Bal-tic Sea [5].

Biofouling is the process by which marine organisms such as algae and barnacles settle on and colonize submerged surface. For boats, biofouling poses a problem as it leads to an increase of the hull surface roughness, thereby impacting both the speed and maneuverability of the vessel through increased drag [6]. Although there are some promising non-toxic fouling prevention methods available for leisure boats, these are not yet well established and the most common way to ward off fouling organisms today is through the use of biocidal AF paints. The main mode of action of an AF paint is to release bio-cides into the water surrounding the hull at a rate such that settling organisms are either repelled or killed through poisoning before they can permanently adhere to the hull [7]. These products, designed to leach toxic substances from the surface of the coating when in contact with water, may however also neg-atively affect non-target organisms. As the use of AF paints typically leads to increased concentrations of biocides in the adjacent water and/or sediments, recreational marinas are specifically prone to contamination [8]. This is prob-lematic as these are commonly located in shallow coastal areas such as shel-tered inlets and coves that are poorly flushed. In the Baltic Sea, these shallow

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inlets are sites of high natural value, important for the reproduction of species and the main habitat for many brackish macro algae [9]. A variety of AF biocides have been used throughout history, most infamously organotin compounds such as tributyltin (TBT), but the vast majority of paints on the market today are copper-based with zinc as a common additive (Chap-ter 2). Although the main function of AF paints is to leach biocides into the water when the boat hull is immersed, there are other pathways by which toxic substances may be emitted into the environment over the lifetime of the paint, such as through hull maintenance activities on land (Chapter 6) (fig. 1).

As AF paints are biocidal products, they are regulated within the EU and require approval from relevant national authorities before they may be mar-keted (Chapter 5). In Sweden, the Swedish Chemicals Agency handles the approval process of AF paints. For a product to be authorized, it must be shown to be efficient against fouling, without posing an unacceptable risk to the environment. Both a paint’s efficacy and its potential for negative effects on the environment is related to its biocidal release rate. The release rate is therefore used to derive Predicted Environmental Concentrations (PECs) in marina waters using the modelling software MAMPEC (Marine Antifoulant Model to Predict Environmental Concentrations), making it a crucial parame-ter of the environmental risk assessment (Chapter 7).

Figure 1. Potential pollution pathways of biocides and substances of concern into

the environment through their use in AF paints on leisure boats.

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2 Antifouling paints

Biocidal antifouling paints are complex products, typically containing 10-15 components, leading to a large diversity in properties between products [10]. Common to all is however that they hold one or more biocides, also called active substances. Ultimately, the environmental risk of an AF paint will be defined by the specific biocides it holds and the speed at which these are re-leased. The latter is in large part be controlled by the paint’s properties.

2.1. Paint types It is not only the type and amount of biocide(s) that will determine the efficacy and toxicity of an AF paint, but also the way in which the active substance(s) is incorporated into the product and consequently released upon use. The paint “binder” is the component that holds the paint matrix together and is often considered to be the most important ingredient [10]. Along with the biocide’s inherent solubility, it is the properties of the binder which will control the manner and speed at which the biocide is released to the water phase. There are currently 3 main biocidal paint types [11]:

• Insoluble matrix paints, also known as contact leaching paints or hard paints, have a binder which will not erode when immersed in seawater (fig. 2A). With time, the soluble biocidal pigments leach from the paint, leaving behind a depleted, porous matrix. As this depleted layer increases in thickness, the diffusion of biocides from the paint film decreases over time.

• Soluble matrix paints, also known as erodible or ablative paints, have a seawater soluble rosin binder. The paint erosion is thus controlled by the dissolution of both the binder and the biocidal pigments (fig. 2B).

• Self-polishing paints are based on acrylic polymers which undergo hydrolysis or ion exchange. The consequent continuous surface re-newal yields a self-smoothing paint surface with a, in theory, steady release of biocides over the lifetime of the paint (fig. 2C).

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Figure 2. Schematics of different paint types, insoluble matrix (A), soluble matrix

(B) and self-polishing (C), and their release of biocides over time.

2.2. Antifouling biocides

Early methods of fouling prevention The most ancient methods of fouling prevention include copper and lead sheathing and tar coatings containing arsenic [7]. Since the 1800’s, however, copper paints have been used as a way to combat fouling on boat hulls. In the 1950’s, antifouling paints mainly contained cuprous oxide (Cu2O), but other biocides such as dichlorodiphenyltrichloroethane (DDT), organomercury and arsenic were also common [12].

Organotin paints In the early 1960’s, the antifouling market was revolutionized with the intro-duction of organotin paints. The organotin compounds (OTCs) tributyltin (TBT) and triphenyltin (TPhT) were the main biocides utilized, firstly in so-called “free association paints”, i.e. mixed in freely into a soluble or insoluble paint matrix [13]. The golden era for organotin paints began however with the introduction of self-polishing type coatings in the 1970’s. In these paints, the

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organotin entity is bound to a polymer through ester bonds which are hydro-lyzed in seawater, leading to a controlled and constant release of the biocide [12]. The self-polishing coatings would ultimately come to be used by > 70% of the world’s fleet, with annual cost savings due to reduced fuel consumption and maintenance estimated to $700.000.000 due to their unprecedented effi-cacy against fouling [14,15]. TBT and TPhT were however later identified as endocrine disruptors, causing imposex, i.e. the development of male sex char-acteristics in the female, in molluscs at very low concentrations [16]. The first reports of environmental effects came at the end of the 1970’s when the col-lapse of oyster farms in Arcachon Bay, France was linked to TBT [17]. As a consequence, France was the first country to ban the use of OTCs in AF paints for recreational vessels in 1982. Organotin paints were forbidden in all EU countries on boats < 25 m in 1989 (Directive 89/677/EEC). As of 2003, the ban was extended to vessels of all sizes (Regulation (EC) No 782/2003).

Copper paints and booster biocides Following the ban of organotin paints, copper formulations (mainly Cu2O and CuCSN) have again become the most widely used [18]. Certain algal species are however tolerant to copper, leading to the addition of organic “booster biocides” to many paints to enhance their performance [6]. In 2001 it was es-timated that 18 different biocidal additives were in use worldwide [19]. Some of the most widely used booster biocides include: irgarol (cubytryne), diuron, seanine (DCOIT), dichlofluanid, chlorothalonil and zinc pyrithione. Although not classified as a biocide in the EU, zinc oxide is a very common additive to antifouling paints for its ability to enhance the erosion process of the coating [20,21].

New biocides There is currently no viable option for widespread replacement of copper in AF paints. The search for new biocidal antifoulants is however actively on-going. New biocide candidates are, for example, the organic substances tri-phenylborane pyridine (TPBP), tralopyril (Econea), capsaicin and medetome-dine (Selektope) [22]. Both tralopyril and medetomedine have already been added to the list of approved biocidal active substances in the EU as of 2015 and 2016, respectively [23].

Non-biocidal methods The growing awareness of the environmental impact of antifouling biocides has led to the development of “greener” fouling prevention methods over the past few decades. These non-biocidal methods typically aim to prevent fouling through physical or mechanical means, rather than chemical ones. The most

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well-known are so-called foul-release coatings which were introduced in the early 1980’s. These consist of silicone or fluoropolymers intended to create a non-stick surface too slippery for fouling to adhere [22]. In 2011, foul-release coatings were estimated to account for around 6% of the total sales by volume for commercial shipping and < 1 % for the recreational market [8,24]. Another green method which has been introduced for leisure boats are stationary boat washes, allowing in-water cleaning of the hull through the action of rotating brushes. There are currently only 24 boat washes along the Swedish coast, but interest in this method is growing [25]. Other options for smaller vessels in-clude the use of hull covers or dry-docking, i.e. storing the boat out of the water when it’s not in use.

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3 Aims

The overall objectives of this PhD thesis was to (1) investigate the pathways through which AF biocides, past and present, may reach the environment through their use on leisure boats and (2) improve the risk assessment of bio-cidal AF paints intended for amateur use. The detailed aims of each paper are presented in Table 1.

Table 1. Aims of the papers included in this thesis.

Paper Aims

I 1. Identify the OTCs present in historic layers on leisure boats; 2. Evaluate whether total tin (Sn) can be used as an indicator for

the presence of banned OTCs on boat hulls.

II

1. Investigate if the soil contamination at recreational boatyards can be linked to hull maintenance activities and whether the age of the boatyard can be related to the extent of contamina-tion;

2. Assess compliance with current AF regulations.

III

1. Investigate the impact of salinity and paint properties on the release rates of Cu and Zn from commercial AF paints;

2. Compare in situ release rates determined with a new X-Ray Fluorescence (XRF) method to release rates derived with cur-rent standardized methods and provided for product authori-zation;

3. Assess whether the products should have gained approval given the measured in situ release rates.

IV

1. Investigate if and how the concentrations and speciation of dissolved Cu and Zn change as a function of boating activity in Baltic Sea marinas and whether water quality guideline val-ues are exceeded;

2. Evaluate whether modelled environmental concentrations (PECs) in MAMPEC are comparable to measured concentra-tions.

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4 Methods

All the papers in this thesis are field studies (papers II and IV) or based on samples either collected from or exposed in the field (papers I and III). A variety of methods, both lab- and field-based, have therefore been utilized throughout this work, as outlined below. More details about the respective methods can be found in the corresponding papers. X-Ray Fluorescence (XRF) analysis has been used extensively throughout this thesis, and the prin-ciples behind this method are therefore initially described in more detail here.

4.1. XRF analysis

Principle XRF analysis is based on the detection and quantification of characteristic flu-orescent X-rays. It is most commonly employed for elemental quantification of solid samples, although applications for liquids also exist. In brief, the non-destructive irradiation of a sample by primary X-rays causes the excitation and consequent ejection of inner shell (K, L) electrons in the atoms. As outer shell electrons transition to fill the created vacancies, i.e. dropping to a lower energy state, the excess energy is emitted as secondary, fluorescent X-rays (fig. 3). As all elements have their own specific set of atomic energy levels, a unique set of secondary X-rays are emitted by each element. These are characteristic to the specific element and the intensity of the emitted X-rays (in counts per sec) will be proportional to the concentration of the element in the sample. Each XRF measurement generates a spectra, allowing the quantification of elements based on peak area [26]. In air, elements as light as phosphorus can be quantified by XRF. A helium atmosphere or vacuum is however required for lighter elements as the low energy of their characteristic X-rays makes them prone to significant attenuation in air [27].

Portable XRF An energy dispersive portable XRF was used in this thesis, allowing both in situ measurements in the field and ex situ measurements in the lab in a stand. Portable XRF analyzers were originally developed to allow for the quick and

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Figure 3. Schematic of a portable XRF and the principle behind XRF analysis.

easy on-site screening of metals in matrices such as alloys and soil. Nowadays, mainly due to the development of better detectors, these handheld instruments can also be used for quantitative measurements [28]. Portable XRFs are typi-cally equipped with X-ray tubes with a voltage between 10 and 50 kV. The instrument used in this work had an X-Ray tube with a voltage capacity of up to 50kV, permitting the excitation and expulsion of the inner, high-energy K-shell electrons of heavier elements such as Sn. The generated characteristic K-lines are generally preferable over the lower order L-lines as their intensity is usually higher (thus yielding lower detection limits) and because they are less prone to overlap with other lines [26].

There are several benefits to portable XRF compared to traditional wet chemical methods: the analysis is direct and non-destructive, allowing for a high throughput and both the start-up and operational costs are low. On the other hand, the detection limit is higher (ppm-range) and although XRF anal-ysis can generate quantitative results, the quality of the measurement is de-pendent on sample properties and the level of sample preparation, as described next [28].

Measurement quality There are several parameters which can affect the measurement and determine the quality of the XRF analysis (screening or quantitative level). The main sources of error are outlined below.

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Physical matrix effects As XRF analysis is performed directly on the solid sample, its physical char-acter is of great importance. As only a small area of the sample is analyzed (Ø 8 mm for the XRF in this thesis), the sample should be as homogeneous as possible in order for the measurement to be representative. Surface morphol-ogy can also influence the measurement if the sample cannot be placed flat against the analyzer window, leading to lower signal intensity. For granular samples, e.g. soils and sediments, particle size is of importance. The smaller the particles, the higher the density of the material present at the analyzed sample surface will be. As water can absorb X-rays, the moisture content of the sample can also influence the measurement [27]. The simplest way to re-duce physical matrix effects for granular materials is therefore to dry and grind the samples prior to analysis. This procedure also serves to homogenize the sample.

Chemical matrix effects Both the primary and secondary X-rays are subject to absorbance from other elements present in the sample. The heavier the elements, the greater their potential to absorb X-Rays. Enhancement effects are also possible whereby one element can be excited by the characteristic X-rays emitted from another. The chemical composition of a sample can therefore affect the intensity of the measured characteristic X-rays. Mathematical models can be used to correct for this kind of matrix effects, but often require in-depth knowledge about the composition of the sample [29]. Another way to avoid biased results is to in-stead ensure that the sample composition as closely as possible resembles that of the standards used to establish the calibration on the instrument. Creating your own matrix-specific calibration is otherwise the best approach to achiev-ing quantitative results, but can be laborsome. Normalizing the measured in-tensities to that of the Compton peak is an additional way to account for matrix effects. The primary X-rays are subject to scattering in the sample and the amount of scattering strongly depends on the sample composition and in par-ticular the amount of light elements. The intensity of the Compton peak re-flects the amount of scattering and, in consequence, the fluorescence absorp-tion of the sample. Normalizing to the Compton peak can thus be likened to normalizing to an internal standard [30].

Spectral interferences Interferences occur when the characteristic X-ray lines of two (or more) ele-ments have energies that are so similar that the detector is not able to resolve them, resulting in peak overlap [26]. Overlap corrections are sometimes pos-sible, but given that all elements have a set of characteristic lines, the easiest solution is often to select another line.

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Analysis of thick or thin samples The sample thickness required for accurate analysis will vary depending on the application. The XRF signal of an element of interest will increase with increased sample thickness until the critical saturation depth (dthick) is attained (fig. 4). dthick thus represents the maximum depth from which secondary X-rays are able to escape from a sample. Beyond this depth, all X-rays are ab-sorbed by the sample and cannot be detected by the instrument. The size of dthick (typically in the µm to cm range) will depend both on the element of interest and the composition of the sample matrix. Higher energy X-rays emit-ted from heavier elements are able to travel longer distances. For the same matrix, heavier elements will therefore be detected at greater depths than lighter ones. The selection of emission line for quantification for an element is also important, as higher order lines (e.g. K) are able to travel longer dis-tances than lower order lines (e.g. L). dthick for the same element will however vary between matrices depending on their composition. As already mentioned, matrices holding larger amounts of heavier elements have a greater potential for absorption, yielding a lower dthick [26,31].

Figure 4. Effect of thickness on the XRF signal. Figure from Ytreberg et al. 2017 [32].

XRF applications are typically divided into two categories based on sample thickness [31]:

• Thick samples: At thicknesses > dthick, the XRF signal is no longer influenced by the sample thickness and the sample is said to be “infi-nitely thick”. For the analysis of bulk materials this is the preferable thickness, as the result is then not dependent on the amount of ana-lyzed sample. Analysis of e.g. soils and sediments, where the result is expressed in mg kg-1, therefore require samples of infinite thickness.

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• Thin samples: At thicknesses < dthick, there is a range where the in-crease in signal intensity is linear due to the absence of absorption or enhancement effects (fig. 4). The cutoff for this window is known as the critical thickness, or dthin. Sample thicknesses <dthin are desirable for the analysis of thin samples, e.g. films or filters. Here, the meas-urement, generally expressed as mass per unit area, is that of the total amount of analyte present in the full thickness of the film.

4.2. Analytical and field methods

XRF applications Three different XRF calibrations were utilized in this thesis (Table 2), two of which had to be developed. In papers I and II, the XRF was used to analyze metal concentrations in bulk samples (paint powder and soil), whereas thin samples (paint films) were measured in paper III. Table 2. Summary of the XRF methods used for metal analysis in various matrices.

Paper Matrix Method

I Antifouling paint powder

Quantification of Sn (in mg kg-1).

II Soil and sediment Quantification of Cu, Zn and Pb (in mg kg-1) using the pre-installed soil calibration pro-vided by the instrument manufacturer.

III Antifouling paint Quantification of Cu and Zn (in µg cm-2) in thin films of AF paint.

In paper I, the XRF instrument was specifically calibrated for the determina-tion of Sn (Kα-line) in ground paint flakes. Paint powder standards were pre-pared by mixing pure Sn powder into a commercially available AF paint for leisure boats, followed by drying, grounding and sieving (<100 µm) into fine powders. The calibration (r2 = 1.000) was then used to quantify the amount of total Sn (in mg kg-1) in ground paint scrapings collected from leisure boats. A measurement time of 30 s was used.

In paper II, metal concentrations (Cu, Zn and Pb) in surface soil were mapped at two boatyards through in situ measurements. Portable XRFs are commonly used for on-site screening of soil at contaminated sites and soil calibrations are readily available for purchase from the instrument retailers [33]. Such a pre-installed soil calibration was used and evaluated here, with a measurement time of 120 s. The soil calibration was also used ex situ in the

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lab stand for the analysis of collected soil and sediment samples which had been dried and ground. Comparison against chemical analysis showed the XRF capable of producing accurate results with a high precision. Finally, the portable XRF was used to in paper III to determine the amount of Cu and Zn (in µg cm-2) in thin films of AF paint based on quantification of their Kα-lines. The first development of the method is described in Ytreberg et al. 2017, where it was determined that the thickness of AF paints films should be < 40 µm to be within the linear range, i.e. to avoid absorbance ef-fects [32]. In paper III, the method was therefore refined: a new calibration was established which better resembled the samples and a paint film applicator was used for all AF coatings (standards and samples) in order to produced paint films with a controlled thickness (< 40 µm in DFT). A measurement time of 30 s was used.

Passive samplers In paper IV, the concentrations and speciation of dissolved Cu, Zn and Ni were measured in two Baltic Sea marinas. The speciation was determined through the use of the passive samplers known as Diffusive Gradients in Thin films (DGT). DGTs accumulate dissolved metals using a chelating gel (Chelex-100 incorporated into a polyacrylamide diffusive gel layer), provid-ing the average in situ concentration of metals in the water during the time of deployment [34]. Conventional discrete water samples only provide the metal concentration (and speciation) at the time of sampling, whereas DGTs provide time-integrated data as they are deployed for a longer period (hours to days) [35]. In paper IV, a deployment time of 3 days was used as it was estimated to be long enough to accumulate a detectable amount of metals, whilst mini-mizing the risk of the device getting covered by fouling capable of blocking diffusion. The gels were eluted in nitric acid and the eluates analyzed by ICP-MS.

Comparisons with more traditional speciation techniques such as analytical voltammetry and ultrafiltration have shown that the DGTs will selectively and reproducibly accumulate free metals, simple inorganic complexes and a con-trolled fraction of small and labile organic complexes [34,36–39]. The DGT-labile concentrations obtained thus reflect the speciation of metals as the de-vices only bind the dissolved species that are labile and considered available to biota [40]. DGTs have been utilized in several previous studies to measure labile concentrations of metals in marinas, ports and shipyards, e.g. to inves-tigate potential correlations between the number of boats and labile metal con-centrations in the water column [41], seasonal variations in labile metals [42] or the impact of tidal-flushing and rainfall events [35].

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Wet chemistry methods

Metal analysis Wet chemistry techniques were employed for the analysis of metals (Cu, Zn, Pb and Sn) in various matrices (Table 3).

Table 3. Summary of the wet chemical methods used for metal analysis.

Paper Matrix Method

I Antifouling paint Acid digestion according to four different methods (see paper). Analysis by ICP-MS.

II Soil and sediment Acid digestion (SS 028150-2), analysis by ICP-OES.

III Antifouling paint Acid digestion, analysis by ICP-MS.

IV Seawater and DGT eluates

Analysis by ICP-MS.

Organotin analysis In paper I, paint samples were scraped from the hulls of 23 leisure boats around the Baltic Sea for the analysis of OTCs and total Sn. A new method for the determination of OTCs in paint flakes had to be developed to ensure maximum recovery of any OTCs contained within the complex matrix of the paint flakes. The paint samples were ground and sieved (<100 µm) into fine powders in order to homogenize the samples and also optimize the extraction recovery.

Current methods for the analysis of OTCs in solid samples such as soil and sediment involve the following sample preparation steps: 1) extraction of the analytes, typically with organic or inorganic acids, 2) derivatization and ex-traction with organic solvents such as hexane or pentane, and 3) pre-concen-tration of the sample [43]. The new method for paint flakes was developed in cooperation with Aarhus University, Denmark, based on their accredited method for sediments [44,45]. Their sediment method involved: ultrasonic ex-traction with HCl and in situ derivatization with sodium tetraethylborate (NaEt4B) followed by phase extraction with pentane. For the analysis of the samples, a gas chromatographer (GC) coupled with a pulsed flame photomet-ric detector (PFPD) was used. PFPD is one of the most common detectors used for OTCs as it has the advantage of reduced interferences from sulphur and phosphorous compounds which can be co-extracted from sediments [43].

The new method for ground paint flakes involved the addition of a pre-extraction step with ultrasonic extraction using a solvent mixture (toluene, ac-etone and propanol), designed to resemble the composition of paint thinner. The procedure for the original sediment method, as described above, was then

20

followed. The new method was able to extract twice the amount of OTCs as the accredited method for sediments. The recovery of OTCs was also com-pared to ISO 23161:2011, a standardized method used for the quantification of OTCs in soil, sediments and sludge and applied by many analytical labor-atories world-wide. This method employs: ultrasonic extraction with a mix-ture of acetic acid:water:methanol (1:1:1), derivatization of an aliquot with NaEt4B and phase extraction with hexane, followed by analysis on GC-ICP-MS. Results showed yet again a notably higher recovery for the new method. The improved recovery of the method developed for paint flakes is likely at-tributed to the added extraction step utilizing mainly non-polar solvents.

4.3. Release rate methods The release rate, i.e. the speed at which toxic substances are released from an AF paint to the water phase, will determine its ability to provide protection against fouling. This parameter, expressed in µg cm-2 d-1, is also the key com-ponent of the environmental risk assessment (see Chapter 7). There are how-ever only a few standardized methods to determine the release rate of AF paints. These methods and their pros and cons are outlined here.

Existing methods

Rotating cylinder method There are currently only two standardized methods accepted by regulatory bodies in EU member states, but concerns have been raised whether these yield representative release rates [46,47]. The first method is a laboratory method known as the “rotating cylinder method” (ASTM D6442-06/ISO 15181:2007) (fig. 5A). Triplicate cylinders are coated with AF paint and im-mersed in holding tanks filled with artificial seawater (pH: 8 ± 0.1, salinity: 33-34 ‰, temperature: 25 ± 1°C). On specific measurement days, the coated cylinders are transferred to cylindrical tanks where they are rotated at 0.2 m/sec for up to 1 hour in 1.5 L artificial seawater [48]. Water samples are collected and chemically analyzed in order to calculate the release rate at the different time points. The final reported release rate is the average rate meas-ured between day 21 and 45. The rotating cylinder method was originally de-veloped to be used by manufacturers as a tool during product development of coatings and not intended to reflect the actual environmental release [47]. The method has also been shown to have poor reproducibility: an inter-laboratory comparison with 6 participating labs yielded coefficients of variation of 24 – 90 % when release rates from 5 different AF paints were compared [49].

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Figure 5. Schematics of different methods to measure release rates from antifouling

paints: the rotating cylinder method (A), the CEPE mass balance equation model (B), the dome method (C) and the XRF method (D).

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CEPE mass balance equation method The other standardized method is a calculation method (ISO 10890:2010) which consists of a mass balance equation developed by the European Council of the Paint, Printing Ink and Artists’ Colours Industry (CEPE). Contrary to the rotating cylinder method, which is both time-consuming and requires costly chemical analyses, this calculation method is quick and inexpensive. The model and release rate equation were derived from the release rate pattern observed for self-polishing organotin and cuprous oxide paints using the ro-tating cylinder method [50]. The model assumes an initial high release rate during the first 14 days, followed by a steady-state release for the remainder of the paint’s service lifetime [51] (fig. 5B). The following equation is used to solve for the steady-state release rate Y (μg cm-2 d-1):

𝑋𝑋 + �𝑌𝑌 × �365 ×𝑡𝑡

12− 14�� = 𝐿𝐿𝐿𝐿 × 𝐿𝐿 × 𝑊𝑊𝐿𝐿 × 100

𝑣𝑣𝑣𝑣 𝜌𝜌 × 𝐷𝐷𝐷𝐷𝐷𝐷 (Equation 1)

o X: amount of biocide released during the first 14 days (μg cm-2) o t: specified lifetime of the paint (months) o La: fraction of the active ingredient in the dry film released during the life-

time t (equal to 0.9 according to European regulatory authorities), o a: mass fraction of active ingredient in the biocide o Wa: concentration of biocide in the wet paint (weight %) o vs: volume solids (%) o ρ: paint density (g cm-3) o DFT: dry film thickness specified for the time t (µm)

Many of the parameters in Equation 1 are fixed and the model does not take binder properties into account as it assumes all paints will leach according to the conceptual model. Thus, the calculated release rate is by default mainly based on the volume of paint applied, the biocidal loading of the paint and the specified lifetime.

Dome method Apart from the standardized methods, another established method is the “Dome method” which was developed by the United States Navy in the 1980’s [52]. Here, ambient natural seawater is recirculated within a polycarbonate dome attached to a submerged ship hull (fig. 5C). Seawater samples for chem-ical analysis are drawn from the enclosed system, allowing the determination of the in situ release rate. Due to the high cost and need for special equipment, this method has been deemed too expensive and impractical to be standardized [47]. The Dome method has however been used to evaluate the other methods. In a study from 2006, release rates for a soluble matrix AF paint derived ac-cording to the two standardized methods were compared to in situ release rates from 5 ships berthed in San Diego (each with different periods of service since

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the coating was applied) [46]. The standardized methods were found to yield higher release rates and correction (reduction) factors of 5.4 and 2.9 were pro-posed for the rotating cylinder and calculation method, respectively. Within the EU, it has since been standard practice that should an AF paint fail the risk assessment using Tier 1 (uncorrected) release rates, it may be repeated with a corrected release rate for a Tier 2 assessment using the proposed factors [53].

The new XRF method Although both the rotating cylinder method and the CEPE calculation method have been found to not be representative, they have been used as an interim solution until a better method for release rate determination is pre-sented [47]. The new XRF method, originally described in Ytreberg et al. 2017 [32] and used in paper III, presents a much better option as it allows the quick and easy determination of in situ release rates. The XRF instru-ment, specifically calibrated for quantitative measurements of Cu and Zn in AF coatings, is used to measure the area concentration (in µg cm-2) of metals on painted panels before and after immersion (fig. 5D). The difference in concentration between two time points represents the amount of metals lost, i.e. the cumulative release. As the immersion time is known, the average re-lease rate over this time period is easily calculated.

The XRF method is quick, with a measurement time of only 30 seconds. Apart from the one-time initial investment of the instrument purchase and its calibration, this method will not require any additional on-going costs within the lifetime of the X-Ray tube (~ up to 6 years, depending on usage intensity). However, contrary to the other methods, the XRF method is by nature re-stricted to the measurement of only metallic or organometallic substances as the analysis is elemental and not compound-specific. AF paints with a mixture of biocides may additionally pose a challenge for this method if the biocides contain the same element e.g. copper pyrithione and Cu2O. The XRF will then not be able to differentiate between the two species and only be able to gener-ate the total and combined metal release from both biocides. The method also requires certain specific sample conditions such as a paint thickness < 40 µm [32]. This is due to the fact that the secondary X-Rays are subject to absorption by metals and these are typically found in large amounts in AF paints. Alt-hough the presence of a thin layer of soft fouling (i.e. slime) was found to not interfere with the measurements in paper III, XRF measurements need to be performed on smooth surfaces. Areas covered with macrofouling such as bar-nacles can therefore not be measured.

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5 Regulation

AF biocides are nowadays regulated at EU level, but specific restrictions and regulations can also exist at national and regional level. Apart from regulations directly concerning AF products, Member States must also comply with envi-ronmental EU regulations with regards to water quality.

5.1. Current regulatory status

Antifouling regulations

In the EU In 2013, the Biocidal Products Regulation (BPR, Regulation (EU) 528/2012) came into force in the EU, replacing the Biocidal Product Directive (BPD, 98/8/EC). According to this new legislation, the approval of active substances now takes place at Union level and biocidal products may only contain ap-proved substances. Authorization of individual products, on the other hand, takes place at Member State level (see Chapter 7). There are currently 10 ap-proved active substances and 2 active substances under review, with authori-zations valid for up to 10 years [23].

How well the BPR succeeds in only approving biocides which will not lead to unacceptable negative effects on the aquatic environment remains to be seen over the coming years, especially with regards to newly approved bio-cides such as e.g. meditomedine, which has already been tagged as an active substance candidate for substitution [54].

Country-specific regulations Even though the BPR seeks to harmonize the regulation of AF paints in the EU, Member States still have the possibility to restrict the use of these prod-ucts at the regional and local level [55]. Both Denmark and Sweden have, for example, banned the use of AF paints in freshwater as the fouling pressure in these waters is low and the environmental risks of using biocidal AF products are deemed to outweigh the benefits [56,57]. The use of AF paints for leisure boats is further regulated regionally in Sweden, with tighter restrictions on the

25

biocidal release from paints used on the East coast (i.e. in the Baltic Sea) com-pared to the West coast, on account of the Baltic Sea being classified as a PSSA [58]. A recent study on the use of AF biocides on German leisure boats showed that although the German recreational fleet was only a third of the Swedish one, 34% more AF biocides were consumed on a yearly basis. The study identified the Swedish regional restrictions as important drivers for re-duced emission of AF biocides to the environment [59]. There seems however to be additional possibilities for reduced emissions in Sweden (and other coun-tries) of e.g. copper from AF paints without compromising the antifouling per-formance. Field studies over three consecutive years (2014-2016) of AF paints authorized in Sweden for use in the Baltic Sea i.e. with lower emissions of Cu, were found to be equally efficient against fouling on the Swedish West coast as paints with roughly twice the emission rate [60]. These results suggest that there is an overuse of copper in current AF paints.

Water quality regulations There are two EU Directives relevant for the assessment and classification of water quality: the Water Framework Directive (WFD) and the Marine Strat-egy Framework Directive (MSFD).

The WFD was adopted in 2000 and aims for the management and protec-tion of European waters (surface water and groundwater) using an integrated river basin approach (2000/60/EC). The main objective of the WFD is for all waters to achieve good chemical and good ecological status. The Directive dictates the division of the waters within each river basin district into units called water bodies. Following the WFD, 33 so-called priority substances (e.g. TBT), presenting a significant risk to or via the aquatic environment, must be monitored in all water bodies (Directives 2008/105/EC & 2013/39/EU). Ad-ditionally, River Basin Specific Pollutants of regional or local importance (e.g. Cu and Zn) should be identified and monitored by Member States. For the evaluation of pollutants, concentrations should be compared to Environmental Quality Standards (EQS). Waters covered under the WFD include all inland waters as well as coastal waters which extend up to 1 nautical mile seaward from the baseline. The chemical status applies even further out, including ter-ritorial waters (12 nautical miles from the baseline).

The MSFD came into force in 2008 and aims at the achievement of Good Environmental Status (GES) in EU’s marine waters by 2020, with focus on their sustainable use (2008/56/EC). The waters covered by the MSFD are all those (starting at the shoreline) within the Member States’ Exclusive Eco-nomic Zone (EEZ). The unit of study equivalent to that of the water bodies under the WFD are here called subdivisions, forming the parts of four distin-guished marine regions: the Mediterranean Sea, the Baltic Sea, the Black Sea or the North-East Atlantic Ocean. The Directive lists 11 descriptors to assess

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the environmental status. For AF biocides, two of these are of relevance: de-scriptors 8 (aim: Concentrations of contaminants give no effects) and 9 (aim: Contaminants in seafood are below safe levels). For Descriptor 8, Member States should monitor the priority substances and specific pollutants already identified under the WFD. Moreover, additional contaminants that may give rise to pollution effects identified through regional or subregional cooperation (e.g. Regional Seas Conventions such as HELCOM and OSPAR) should be assessed (Commission Decision (EU) 2017/848). Thus, there is overlap be-tween the WFD and MSFD, both geographically (coastal waters), and in terms of substances to be monitored and assessed against established EQS-values.

5.2. Regulatory challenges

Introductions and bans – a recurrent pattern As mentioned in Chapter 2, the ban of OTCs in AF paints for amateur use in 1989 led to the return to copper paints, commonly with the addition of booster biocides. However, it only took a few years before elevated concentrations of many booster biocides were reported in both water and sediments across Eu-rope [61]. Restrictions for use on small vessels were soon to follow for several of these substances in individual EU countries [62] (fig. 6). Some biocides such as TBT, diuron and irgarol have also been identified as priority sub-stances at EU level (Directives 2008/105/EC & 2013/39/EU). TBT is addi-tionally tagged as a Priority Hazardous Substance, signifying that its emission should be actively phased out by Member States. The introduction of new bi-ocides onto the recreational antifouling market, typically followed by their ban when they are discovered to be harmful to non-target organisms seems to be a recurrent pattern for AF biocides [62].

The problem of historic paint layers Restrictions on specific substances such as some of the booster biocides men-tioned in Chapter 2 have been found to lead to reduced concentrations in the environment [63]. However, as many of these compounds are persistent they can still be detected in the environment long after their ban. This is particularly true of OTCs such as TBT which degrade very slowly in sediments where they can persist for decades [64]. Additionally, historic paint layers on leisure boats can be problematic. Even though a substance is restricted, old paint layers may not necessarily be stripped from the hulls. More common practice is that the paint is overcoated with new paint layers. Overcoating may partially or com-pletely block the emission from the underlying layers, but only as long as the

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integrity of the barrier layer(s) is not compromised. The complete phasing-out of a banned substance may therefore be a lengthy process.

Several studies suggest that OTCs are still emitted to European waters from leisure boats, likely due to their presence in historic paint layers [65–69]. In paper I, OTCs were detected in all paint scrapings collected from 23 leisure boats in Sweden, Finland and Germany. The leisure boats were deemed to have been built prior to 1989, suggesting that the OTCs were likely located in old paint layers and not the result of recent illegal paint use. When it comes to regulation and monitoring, TBT has received a lot more attention than TPhT, despite the fact that both substances have similar toxic properties [70]. This is probably due to the fact that TBT has been perceived to have been more widely used in AF paints. The results from paper I showed however that both compounds were commonly detected on the sampled leisure boats. The study also revealed difficulties for some current soil and sediment methods regard-ing the extraction of particularly TPhT. This may have led to underestimation of environmental concentrations, contributing to the perception that TPhT was not used as much as TBT.

Most importantly, paper I showed a clear linear relationship between the total Sn concentrations measured in the paint scrapings and the amount of Sn present in the form of OTCs. Total Sn, which is much easier and cheaper to measure than OTCs, is therefore a viable indicator for the presence of OTCs. A method to screen for Sn on boat hulls by means of portable XRF has been introduced in recent years [71,72]. The result from paper I confirms that this is a valid method to identify vessels bearing layers of banned OTCs and there-fore in need of remediation in order to reduce the emission of OTCs to the environment. This particular screening method is however only made possible by the fact that OTCs are organometallic substances. For other banned sub-stances which do not contain metals, the development of screening methods becomes more difficult. Measurements by portable XRF have now shown that roughly 10 % of ships and 25 % of leisure boats inspected in Sweden hold Sn concentrations indicative of banned organotin paint [73].

Compliance issues Apart from historic layers, continued emissions to the environment of a banned antifouling agent can also be the result of illegal use. Bans on AF bi-ocides have historically tended to firstly target paints for amateur use with bans for commercial vessels not introduced until years later. OTCs were, for example, firstly banned on leisure boats in 1989 in the EU, but not banned on ships until 2003. Additionally, the world-wide ban through IMO did not come into effect until 2008 under the Antifouling Systems (AFS) Convention [74].

The use of ship paints on recreational vessels is one of the most common forms of non-compliance, along with the use of imported paints and the use of paints not compatible with specific regional restrictions. A national survey

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for leisure boat owners conducted in Sweden in 2010 showed that 2.7 % of boat owners painted their boats with ship paints, whilst another 2.5 % report-edly used imported paints. The number of unreported instances of non-com-pliances may however be larger as an astounding 36.7 % reported to not know what their boat was painted with [75]. In 2015, the follow-up survey addressed the issue of non-compliance with regional restrictions and found, for example, that 12 % of boat owners moored in lakes were using paints that were not compatible with current regional legislation, i.e. the ban of biocidal AF paints in freshwater introduced in 1992 [56,76]. In paper II, the illicit use of biocidal AF paints at a Swedish boatyard by lake Mälaren, Stockholm was made evi-dent through the elevated Cu concentrations detected in both surface sedi-ments and the top soil of younger areas (used for < 10 years) of the investi-gated boatyard.

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6 Pollution pathways

Antifouling biocides will be released from the paint into the water phase upon immersion of the coated boat hull. There are however other pathways by which the biocides and other substances of concern in the paint will reach not only the aquatic environment but also the terrestrial one.

6.1. In the harbor

In-water leaching Leisure boats are typically moored in their home harbor during the majority of the boating season. A survey for Swedish boat owners in 2015 showed, for example, that leisure boats were only actively used during 16 days i.e. 10 % of the boating season [76]. Recreational harbors are thus especially prone to accumulation of AF biocides in both water and sediments, as illustrated by various studies of marinas in Europe and elsewhere [41,42,67,59,77–81]. This was also the case for the harbors studied in this thesis. In paper II, surface sediment concentrations in a harbor by lake Mälaren, Stockholm, Sweden, ranged between 70-120 mg Cu/kg and 230-320 mg Zn/kg, thus typically ex-ceeding derived PNEC-values for freshwater sediments of 87 mg Cu/kg and 179 mg Zn/kg [82,83]. An increase in concentration with decreased depth was also measured, suggesting an increased use of AF paints with time despite the harbor’s location in freshwater where the use of biocidal AF paints has been restricted since 1992 (see 5.1.).

The concentration of total dissolved Cu was studied in two marinas by the Baltic Sea over an entire boating season in paper IV. The results show roughly a doubling of the concentration during peak boating season, leading to exceedance of the Predicted No Effect Concentration (PNEC, see Chapter 7) (fig. 7A). Both Cu and Zn were also found to exceed the Swedish EQS-values at least two-fold. The marina waters therefore fail to attain “Good sta-tus”, as specified under the WFD. Comparison with concentrations measured in other Baltic Sea marinas in Sweden and Finland (Table 3 in paper IV) indicates that the findings are not unique. Analysis based on a geographical survey from 2008 showed that there are on average 4.3 marinas and 151 jetties per water body in Swedish coastal waters. For some water bodies, as many as

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Figure 7. Total dissolved and DGT-labile Cu concentrations (A) and proportion of

DGT-labile Cu concentrations (B) measured in 2016 in one of the studied Baltic Sea marinas in Paper IV (Bullandö Marina, Stockholm, Sweden). The red triangle shows

the Swedish PNEC-value for total dissolved Cu (1.45 µg/L) for the Baltic Sea.

Figure 8. Aerial photo showing a water body (blue outline) in the Stockholm archi-pelago holding 23 marinas (red symbol) and 833 jetties (yellow symbol) based on a

survey from 2008 [84].

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23 marinas may however be present (fig. 8). The use of biocidal AF paints therefore has the potential to impact the status of many water bodies.

The concentration of bioavailable Cu was also monitored in the marina through the use of passive samplers (DGTs). The results showed that the con-centration of DGT-labile Cu follows the same time trend as total dissolved Cu. However, the proportion of total dissolved Cu which can be considered bioavailable was found to double from ~ 20 to 40 % during peak season (fig. 7B). The concentration of dissolved Cu was thus not only significantly in-creased, but the Cu speciation in the marina waters was also altered in favor of bioavailable species. A similar result was observed for Zn (see paper IV). The change in speciation is likely caused by the release of Cu and Zn in bioa-vailable form from AF paints.

Boat washers Boat washers may act as additional sources of biocides to the marina waters. Although intended to be used to clean unpainted hulls from fouling, some Swedish municipalities have introduced a temporary exception to this rule in order to promote the use of boat washers. All boat washers in Sweden are equipped with a sedimentation basin in order to collect the generated waste, but a recent study found that these may not be efficiently capturing all gener-ated paint particles [85]. The question is whether the long-term goal of getting boat owners to use more sustainable fouling prevention methods outweighs the short-term environmental hazard caused by the in-water brushing of hulls still coated with biocidal AF paint. In paper I, very high concentrations of TBT and TPhT were detected in a sediment sample collected from a boat washer basin. The mechanical brushing of boat hulls is thus also causing the release of biocides contained within historical paint layers, suggesting some caution may be needed.

6.2. In the boatyard Boatyards are used for storage of leisure boats during the winter time in Baltic Sea countries. They are also the sites where hull maintenance activities occur before and after the boating season. In the springtime, boats are prepared for the upcoming season. This mainly entails scraping and sanding of the hull followed by recoating with new layers of AF paint. In the fall, boats are re-trieved from the water and the hull is cleaned from fouling, typically using high pressure hosing. Some larger marinas may have dedicated wash pads with water treatment where the pressure hosing takes place. Pressure hosing may otherwise take place directly over the slipway, with the waste water emit-ted into the marina waters.

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Most of the hull maintenance tends however to be carried out over the ground intended for the storage of the individual boat. If the soil is not cov-ered, paint particles generated from the maintenance end up on the ground where they will leach the biocides they hold, polluting the soil [86]. More persistent AF paint particles may also be transported with runoff and wind into nearby marina waters and settle in the sediments. Here, the particles can leach their biocides through dissolution or abrasion, and also be ingested by benthic organisms [87]. The key finding in paper II was the significant difference in top soil concentrations of Cu and Zn based on land use detected in the two investigated boatyards. Areas dedicated to boat storage were found to hold significantly higher concentrations of the two metals compared to areas used for other purposes (e.g. roads and car parking). Concentrations as high as 226 and 25 times the national Swedish guideline values for industrial land for Cu and Zn, respectively, could be measured. Pb, which has been used historically, both as the primary toxicant and for pigmentation purposes in AF paints [88], was also measured in paper II. Some elevated concentrations were occasion-ally detected in the oldest boat storage areas (> 50 years old). None of the soil measurements in the areas used for < 10 years were however found to exceed the guideline value. The active phasing-out of Pb substances from AF paints and other consumer products since the mid 1990’s therefore seems to have had effect [89]. Other historic biocides, such as TBT, may however be more ubiquitous on boat hulls (see 5.2.) and caution should therefore be exercised during hull maintenance.

Without containment of the generated waste, sanding and hosing can pro-mote the emission of both present and past biocides into the environment. Due to the high concentrations of metals in AF paints, measurable pollution of boatyard soil may only take a few years. The surface soil in areas used for only 9 years in one of the investigated boatyards in paper II was already hold-ing significantly elevated concentrations of Cu and Zn. Given that this boatyard was located by freshwater where the use of copper paints had been banned for nearly 25 years, it may take even less time for boatyards located along the Swedish coast to become polluted.

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7 Risk assessment

To ensure that the use of AF paints does not negatively impact the environ-ment and jeopardize set environmental goals, environmental risk assessments are performed. The work carried out in both papers III and IV related to the exposure assessment from AF paints and how this part of the risk assessment procedure could be refined.

7.1. Environmental risk assessment procedure

Current procedure Although the active substances in AF products are authorized at EU level un-der the BPR, Member States are in charge of the authorization of individual products before their placing on the national market. To be authorized, a prod-uct must fulfil two requirements: be effective against fouling whilst not posing unacceptable risks to the environment. The latter is evaluated through an en-vironmental risk assessment. In the EU, the potential environmental risk posed by an AF paint is evaluated from two parts: the hazard posed by the active substance(s) in the product and the potential environmental exposure (fig. 9) [90].

The hazard associated with the substance is related to its potential for toxic effects on aquatic organisms and is represented by the Predicted No Effect Concentration (PNEC). The PNEC denotes the highest concentration of the substance at which no toxic effects are observed. The exposure assessment, on the other hand, is represented by the Predicted Environmental Concentra-tion (PEC), i.e. the concentration in the environment that the use of the product is estimated to yield. The main focus of the environmental exposure is the passive leaching from AF paints to marina waters. Hence, the release rate to the water phase (expressed in μg cm-2 d-1) of both active substances and sub-stances of concern contained within an AF paint is a fundamental parameter of the risk assessment. The release rate is used to model the PEC in marina scenarios in the software MAMPEC. In MAMPEC, EU member states may utilize custom national marina scenarios and two such scenarios exist for Swe-dish waters: an East coast (Baltic Sea) and a West Coast scenario [58].

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Figure 9. Environmental risk assessment scheme for antifouling paints in the EU.

For the determination of the ultimate risk associated with an AF paint, a risk characterization ratio (RCR) is determined. Here, the PEC is assessed against the PNEC: RCR = PEC / PNEC. To take into account mixture effects, the Swedish Chemicals Agency employs the Concentration Addition approach according to which the RCRs of all active substances and substances of con-cern are summed together. The calculated summed RCR is then compared to “1”, with RCR > 1 indicating an unacceptable risk.

New procedures As of 2018, there are some new practices in place for the authorization of AF paints in the EU. The aim is to facilitate the concept of mutual recognition under the BPR, whereby the authorization of a product in one Member State can be extended to another. Mutual recognition has so far has not been used extensively due to differences in risk assessment procedures between Member States [55]. The new procedures aim to harmonize the risk assessment by cat-egorizing 148 national MAMPEC marina scenarios in marine waters by re-gion: Atlantic, Mediterranean, Baltic Transition or Baltic Sea. The PEC of the biocide(s) for a given product will be calculated for all marinas within a re-gion, with the 90th percentile concentration intended to represent a realistic worst case and form the basis for the calculation of the RCR [90].

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7.2. Towards an improved risk assessment

Improved release rate estimates

XRF method vs existing standardized methods In paper III, Cu and Zn release rates for five commercially available AF paints for amateur use, containing Cu2O and ZnO at varying concentrations, were determined by XRF. Panels coated with the paints were immersed up to 84 days at two Swedish marinas with different salinities (5 and 14 PSU). The release rates determined with the XRF method were compared to those used for the product authorization. The latter were generated by means of either of the two standardized methods, i.e. the rotating cylinder method or the CEPE mass balance equation (see 4.3.). When comparing to the Tier 1 (uncorrected) release rates, it is clear that the standardized methods can both overestimate or underestimate the release rate, all depending on paint and salinity (fig. 10). For almost all the paints, the default correction (reduction) factors for the ro-tating cylinder and the CEPE mass balance equation methods had been applied in order to pass the risk assessment and gain product authorization. The result-ing Tier 2 release rates were found to underestimate the release of both Cu and Zn several-fold compared to the XRF release rates measured at salinities cor-responding to those of the intended use. The use of the default correction fac-tors were therefore deemed inappropriate for Baltic Sea waters.

Figure 10. Average release rates across 84 days for Cu (A) and Zn (B) for the 5

studied paints (paint A – E) in paper III. The release rates were determined with the XRF method at two different salinities. Error bars show the standard deviation of 4 replicate panels. For the authorization of the paints, standardized methods in the form of the rotating cylinder method (R) and the CEPE mass balance equation method (C) were used. Shown here are their Tier 1 (uncorrected) release rates.

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Parameters affecting the release Several parameters, both environmental and connected to the paint’s formula-tion were identified as important for the release of Cu and Zn in paper III and should therefore be considered when determining relevant release rates to be used for the risk assessment of AF paints.

Environmental conditions In paper III, the leaching of especially Cu was found to be significantly af-fected by salinity, with twice as much Cu released at 14 PSU compared to 5 PSU over a period of 84 days (fig. 10A). Although to a lesser extent, salinity was also shown to have a significant effect on the release of Zn for three of the paints. Salinity is thus a very important parameter which should be taken into account when determining a paint’s release rate. Other parameters such as temperature and pH have also been shown to influence the leaching of AF paints in previous studies [91–94]. The use of relevant environmental condi-tions when determining the release rate is therefore especially important for the Baltic Sea whose salinity and temperature range will differ quite substan-tially from those of the standardized methods (33-34 PSU, 24-26°C).

Although some natural seawater conditions (e.g. salinity, temperature, pH) can be mimicked in the laboratory, others are more difficult to replicate. Bio-films may, for example, affect the release negatively through binding of the released metals and/or positively by changing the conditions (e.g. pH) at the coating surface [95,96]. By exposing the painted panels to natural seawater in the field, any potential effect on the biocidal release due to the formation of a biofilm on the surface of the coating will be accounted for. Additionally, the immersion of panels in marinas ensures that the coatings are exposed to rele-vant hydrodynamic conditions. Although the Dome method is also able to de-termine in situ release rates, it has never been proven that the hydrodynamic conditions inside the dome actually reflect those outside [96]. The results from paper III, in particular the effect of salinity, emphasize the importance of de-termining release rates under field conditions which correspond to those of the intended use of the product.

Paint and substance properties The change in release rate over time was also examined in paper III. The paints included in the study were of two different types: 4 paints were so-called soluble matrix paints and 1 paint was of a self-polishing kind (see Chap-ter 2). The results revealed that the change in release rate behavior with time was paint-specific. In other words, even though 4 of the paints were of the same “type”, their release rate behavior differed. Cu and Zn were also found to leach in different ways, most likely due to the fact that Cu2O is more soluble in seawater than ZnO [20]. These findings reveal the unsuitability of applying a uniform release rate prediction model such as the CEPE calculation method

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to all AF paints and substances. Indeed, if one compares the change in release rate with time as modelled from the XRF measured release rates in the two marinas to the theoretical CEPE model, it is clear that the latter does not pro-vide a realistic representation of the leaching of Cu and Zn from AF paints in Baltic Sea marinas (fig. 11). The two paints exemplified in figure 11 show that the (Tier 1) CEPE model can lead to both underestimation and overestimation of the environmental release rate, all depending on the paint, the substance and the salinity. In this sense, the rotating cylinder method is preferable as the release rates generated should, to some degree, reflect paint- and substance-specific behavior. This laboratory method is however still limited by the fact that it is not intended to replicate the leaching that will occur during natural, environmental conditions.

Figure 11. Release rates at 5 and 14 PSU modelled from the XRF release rates

measured in marinas (from paper III), to be compared to the CEPE mass balance model. Also shown are the Tier 2 CEPE release rates, equal to the Tier 1 steady-state release rate divided by the correction factor of 2.9. Exemplified here are: a soluble matrix paint (paint B from paper III) and a self-polishing paint (paint E).

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Improved modelling of environmental exposure In order to obtain realistic PECs in MAMPEC, two criteria must be fulfilled: (1) the release rate used as input into MAMPEC must be representative and (2) the MAMPEC model parameters must be correct. These two criteria were investigated in papers III and IV as detailed below.

Importance of modelling with representative release rates As already discussed, the current standardized methods and the common use of correction factors typically lead to underestimation of the release rate. For the five studied AF paints in paper III, the release rates of Cu and Zn were underestimated up to 8-fold and 3-fold respectively using standardized meth-ods. The release rates that were instead determined with the new XRF method were used to re-calculate the RCRs for each of the paints. Concentrations were modelled for the national Swedish marina scenarios and version 2.5 of the MAMPEC software was used as that was the requirement at the time of the products’ approval [53]. The calculated RCRs were found to be between 1.5 and 4, indicating that none of the studied paints would have been allowed on the market if realistic release rates had been used for the risk assessment.

Importance of accurate model parameters In paper IV, dissolved concentrations of Cu and Zn were measured across a boating season in the same marina upon which the Swedish East Coast (Baltic Sea) MAMPEC marina scenario is tailored (fig. 7A). The observed concen-trations were compared to the modelled PECs for the three AF paints author-ized for use in the Baltic Sea that were studied in paper III. The PECs were derived using release rates derived both from standardized methods (as pro-vided by the manufacturers) and with the new XRF method. For Cu, only the PECs derived from the XRF release rates were high enough to explain the observed marina concentrations. This emphasizes again the importance of us-ing realistic release rates for the derivation of PECs. Although the PECs for Zn derived with the XRF release rates improved the comparability to the ob-served concentrations, these were still not high enough to account for the lev-els of dissolved Zn detected during peak boating season. The main underlying cause for the discrepancy was found to be the use of an inappropriate parti-tioning coefficient between dissolved and particulate phases (KD) in the mod-elling of the PECs. Using a revised value for KD did however result in more realistic PECs. Paper IV thus concludes that representative PECs can be gen-erated in MAMPEC if the input data (i.e. the release rate) is realistic and if the model parameters are correct. The paper highlights the importance of having model marinas in MAMPEC tailored after real marinas so that validation of model results can be made possible.

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Speciation The speciation of dissolved metal species is currently not taken into consider-ation when assessing the environmental risk of AF paints. Only the total dis-solved concentrations are evaluated. The results from paper IV show however that the use of AF paints significantly altered the speciation of both Cu and Zn in marina waters, with a doubling of the fraction of dissolved metals which may be considered bioavailable (fig. 7B). More research is however needed in order to assess whether speciation should be considered for a more refined risk assessment.

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8 Conclusions and future perspectives

This thesis highlights some of the challenges facing biocidal AF technologies with regards to sustainability. The intrinsic problem of AF paint biocides is their need to be toxic against a broad range of fouling organisms at their point of release at the hull, without being toxic to non-target organisms. An ideal biocide should therefore have a broad spectrum activity, combined with low mammalian toxicity, low bioaccumulation rate and a high degradation rate in the marine environment [50]. No biocide which can fulfill all of these require-ments is currently marketed. As shown in this thesis, leisure boats can remain sources of banned biocides to the aquatic environment long after the introduc-tion of restrictions, due to historic paint layers. Banned AF biocides may also linger in soil and sediment for a long time. Hence, there is a need for caution when authorizing new substances. Although regulations are key to reducing the environmental hazard posed by AF paints, this thesis also highlights the need to ensure compliance with current legislation both when it comes to boat owners’ hull maintenance practices and their choice of AF paint. Some sub-stantial short-comings in the current environmental risk assessment procedure for AF paints were also revealed through the work presented here. For one, the current standardized methods to estimate the biocidal release rate were found to be lacking. Consequently, products that in reality pause an unaccepta-ble risk to the environment have been approved, causing exceedance of water quality guideline values in marinas. The findings here show the need for paint-specific release rates determined under conditions reflecting the intended use of the product. As the market is currently dominated by copper-based paints, the new XRF method is an excellent alternative that holds great potential for standardization. Secondly, it is crucial to ensure that the MAMPEC model yields realistic PECs. It is therefore of immense value to have model scenarios tailored after real-life marinas so that the model results can be validated. Here, it allowed for the pin-pointing of the specific parameters which need address-ing in order to model PECs consistent with observed concentrations. Several new methods have been developed as part of this thesis. The new method for OTCs in paint flakes revealed that current methods may greatly underestimate the amount of organotins in soil and sediment samples contain-ing paint flakes. Potential environmental risks related to OTCs could currently

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be underestimated or overlooked due to insufficient extraction of these com-pounds from the sample. This is of concern as TBT is a Priority Hazardous Substance currently being monitored in sediments in EU waters. The new method for determining release rates is unique and at the forefront when it comes to the research and development of new XRF applications. Hopefully, it can be utilized by society, not only to improve the risk assessment of AF paints, but also for the development of more environmentally friendly paints. With this new technique, the tailoring of AF paints efficient against fouling but with a minimum release of Cu and Zn is made possible. Additionally, through better estimates of the release rates, the loads of Cu and Zn from lei-sure boating and shipping to the coastal water ecosystems can be better as-sessed, aiding the sustainable management of these activities. Moving forward, the consequences of the new risk assessment procedures in-troduced in 2018 will have to be seen. The principle of mutual recognition will likely mean that the number of AF products on the market will increase. Mon-itoring in the marine environment of the newly authorized biocides under the BPR, such as medetomedine and tralopyril, should be carried out in order to be able to detect potential signs of environmental hazard at an early stage.

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Acknowledgements

Firstly, I have to thank my awesome supervisors without whom this thesis would not exist. Ann-Kristin, big thanks for all your support and for being so dependable. Remote supervising is perhaps not the best setup for a PhD but we managed to make it work somehow! It has been reassuring to know that you were just a phone call away and that I could always count on you. Erik, we have been through some rough times but it all worked out in the end, I think. Thank you for all the fun and challenging scientific discussions. I hope you’ve enjoyed them as much as I have! Britta, thank you for your enthusi-astic support during the first half of this PhD journey. Thank you, Kent, for checking up on me and for listening and providing sensible advice. This PhD has provided me with two academic families, and so I would like to thank both my colleagues at ACESx (SU) and MES (Chalmers). A special thanks to Maria B and João for being awesome coworkers as well as friends. Karin A and Karin S, thanks for all the administrative support and for always being so friendly and helpful. Per I, thank you for all your (crucial) help with all things IT. Dinis, thanks for being such a great office mate! A special thank you to Josefin for the pep talks and wonderful company during lunches, fika breaks and AWs over these last years. Thank you, Gunnar, for all your support and calming words during times of crisis. Markus, your help in reviewing both my licentiate and doctoral theses has been greatly appreciated! Thanks to my family and friends. Special thanks to Jan for his help with the printing of this thesis. Marie, thank you for always being there, even though we seldom find ourselves in the same country (or time zone). Andreas, there are not enough words to express how grateful I am for your love and support over these past few years. You’ve been there for both the highs and lows (and the low lows) on this wild ride that took us from the US back home to Stockholm and, later on, down to Gothenburg. I couldn’t have done this without you and I hope I can be an equally supportive partner to you during your PhD journey. Iris, thank you for the (literal) motivational kicks during the writing of this thesis. ♥