Unraveling the importance of solid and

adsorbed phase mercury speciation for

methylmercury formation, evasion and

bioaccumulation

Sofi Jonsson

Department of Chemistry

Umeå 2013

© Sofi Jonsson, Umeå 2013

This work is protected by the Swedish Copyright Legislation (Act 1960:729)

ISBN: 978-91-7459-546-8

Cover: Sofi Jonsson (Fron cover: perch from Hennan, Ramsjö. Back cover: red (α-

HgS(s)) and black (β-HgS(s)) cinnbar, Mesocosms, Umeå Marine Sciences Centre)

Electronic version available at http://umu.diva-portal.org/

Printed by: the service centre in KBC, Umeå University, January 2013

Umeå, Sweden 2013

Till Mathias ‒ min älskade och saknade

i

Table of Contents

List of publications ii

Abstract iii

Abbreviations v

Sammanfattning (Summary in Swedish) vi

1. Introduction 1

2. Experimental systems and approaches 8

2.1 Hg speciation and fractionation analysis in environmental samples 8

1.1 Application of Hg stable isotope experimental approaches 10

1.1.1 Determination of gaseous Hg0, HgII, MeHg and DMHg 12

1.1.2 Methylation and demethylation rates and rate constants 14

1.1.3 Determination of methylation rate constants using solid or adsorbed

chemical forms of HgII (paper II) 18

1.2 Scale of experiment – from micro- to mesocosms 19

3. Mercury methylation, evasion and bioaccumulation 21

3.1 Methylation of HgII 22

3.2 Demethylation of MeHg 26

3.3 Evasion of Hg 27

3.4 Bioaccumulation of mercury 28

4. Implication and final remarks 35

Acknowledgements 41

References 44

ii

List of publications

Publications included in this thesis:

I. Sofi Jonsson, Ulf Skyllberg, Erik Björn (2010), Substantial

Emission of Gaseous Monomethylmercury from Contaminated

Water-Sediment Microcosms, Environmental Science and

Technology, 44, 278-283

II. Sofi Jonsson, Ulf Skyllberg, Mats B. Nilsson, Per-Olof

Westlund, Andrey Shchukarev, Erik Lundberg, Erik Björn

(2012), Mercury Methylation Rates for Geochemically Relevant

HgII Species in Sediments, Environmental Science and

Technology, 46, 11653-11659

III. Sofi Jonsson, Ulf Skyllberg, Mats B. Nilsson, Agneta

Andersson, Erik Lundberg, Erik Björn, Differentiated reactivity

of geochemical mercury pools control methylmercury levels in

sediment and biota, Manuscript

IV. Sofi Jonsson, Agneta Andersson, Mats B. Nilsson, Ulf

Skyllberg, Erik Lundberg, Jeffra K. Schaefer, Staffan Åkerblom,

Erik Björn, Impact of Nutrient and Humic Matter Loadings on

Methylmercury Formation and Bioaccumulation in Estuarine

Ecosystems, Manuscript

Paper I and II is reprinted with permission from publisher. Copyright © 2010

and 2012, American Chemical Society.

iii

Abstract

Monomethylmercury, MeHg, is formed under anoxic conditions in waters,

sediments and soils and then bioaccumulated and biomagnified in aquatic

food webs, negatively effecting both human and wildlife health. It is

generally accepted that precipitation of mercury, Hg, and adsorption of Hg

to e.g. organic matter and mineral surfaces are important processes limiting

the reactivity of Hg mobilized in the environment by natural and

anthropogenic activities. However, knowledge concerning the role of

different solid and adsorbed chemical forms of Hg for MeHg formation,

evasion and bioaccumulation is missing. Such information is vital for the

understanding of environmental processes controlling MeHg formation and

bioaccumulation, as well as for predicting how changes in e.g. loading rates

of atmospheric Hg and the outcome of climate change scenarios and

anthropogenic land use could alter Hg concentrations in biota.

In this thesis, a novel experimental approach, using isotopically enriched

solid and adsorbed phases of inorganic Hg, HgII, as tracers, was developed.

Using this approach, we successfully determined rates of MeHg formation

from solid and adsorbed Hg species in sediment slurries and in mesocosm

systems under conditions closely resembling those in field. We conclude that

the solid/adsorbed phase speciation of HgII is a major controlling factor for

MeHg net formation rates. Microcosm experiments revealed that newly

formed MeHg was a major contributor to the evasion of MeHg from the

water‒sediment system, emphasizing the importance of MeHg formation

rate, rather than MeHg concentration, in the sediment for this process. From

mesocosm systems, we provide experimental evidence, as well as quantitate

data, for that terrestrial and atmospheric sources of HgII and MeHg are more

available for methylation and bioaccumulation processes than HgII and

MeHg stored and formed in sediments. This suggests that the contribution

from terrestrial and atmospheric sources to the accumulation of Hg in fish

may have been underestimated. As a consequence, in regions where climate

change is expected to further increase land runoff, terrestrial MeHg sources

may have even higher negative effects on biota than previously thought.

Data and concepts presented in this thesis lay the basis for unprecedented in-

depth modeling of processes in the Hg biogeochemical cycle that will

iv

improve our understanding and the predicting power on how aquatic

ecosystems may respond to environmental changes or differences in loading

rates for atmospheric Hg.

v

Abbreviations

BAF Biota Accumulation Factor

BSAF Biota-Sediment Accumulation Factor

DMeHg Dimethylmercury

DOM Dissolved Organic Matter

d.w. Dry Weight

EXAFS Extended X-ray Absorption Fine Structure

GC Gas Chromatography

GEM Gaseous Elemental Mercury

HPLC High Performance Liquid Chromatography

HgII Inorganic divalent mercury

ICPMS Inductively Coupled Plasma Mass Spectrometry

IDA Isotope Dilution Analysis

IUPAC International Union of Pure and Applied Chemistry

KD Solid-liquid partition coefficient

kd Demethylation rate constant

km Methylation rate constant

LOD Limit of Detection

MeHg Monomethylmercury

METALLICUS Mercury Experiment to Assess Atmospheric Loading in

Canada and the United States

NOM Natural Organic Matter

rd Rate of demethylation

RGM Reactive Gaseous Mercury

rm Rate of methylation

Sed Sediment

UNEP United Nations Environmental Program

US EPA United States Environmental Protection Agency

Wt Water

XAS X-ray Absorption Spectroscopy

XANES X-ray Absorption near-edge Structure

vi

Sammanfattning (Summary in Swedish)

Monometylkvicksilver, MeHg, bildas under anoxiska förhållanden i

naturliga vatten, sediment och jordar och bioackumuleras och magnifieras

därefter i den akvatiska näringskedjan med negativa effekter på djur och

människor som följd. Det är generellt vedertaget att utfällning av Hg och

adsorption av Hg till exempelvis organiskt material och mineralytor

begränsar tillgängligheten för biogeokemiska reaktioner av Hg som

mobiliserats i miljön via naturliga och antropogena processer. Kunskap om

betydelsen av speciationen av Hg i fasta och adsorberade faser för bildning,

avgång och bioackumulering av MeHg är dock bristfällig. Denna

information är kritisk för att förstå vilka processer som kontrollerar bildning

och bioackumulering av MeHg samt för att kunna prediktera hur olika

ekosystem kan förväntas svara på exempelvis ändrad deposition av

atmosfäriskt Hg eller hur klimatförändringar kan påverka koncentrationerna

av Hg i fisk.

I denna avhandling har en experimentell metod utvecklades, där

isotopanrikade fasta och adsorberade kemiska former av oorganiska tvåvärt

Hg, HgII används som s.k. ”tracers”. Denna metod användes för att

bestämma MeHg bildningshasigheter i homogeniserade sediment prover

samt i mesokosmsystem där förhållandena efterliknar de som förväntas i

naturliga ekosystem. Från dessa drar vi slutsatsen att speciationen av HgII i

fast/adsorberad fas är en viktig kontrollerande faktor som begränsar

nettobildningen av MeHg. Mikrokosmexperiment visade att i första hand

nyligt bildad MeHg avgick till gasfas vilket understryker betydelsen av

MeHg bildningshastighet, snarare än koncentration, i sedimentet för denna

process. Från mesokosmexperimenten visar vi, med kvantitativa data, att

terrestra och atmosfäriska källor av HgII och MeHg är mer tillgängliga för

bildning och bioackumulering av MeHg än HgII och MeHg lagrat eller bildat

i sedimenten. Orsaken till detta är framförallt skillnad i speciationen av Hg i

fasta/adsorberade faser. Detta innebär att bidraget från MeHg från terrestra

och atmosfäriska källor till koncentrationen av Hg i fisk kan ha

underskattats, samt att de negativa effekterna på MeHg exponering i

områden där exempelvis klimat-förändringar förväntas leda till ökad terrest

avrinning kan bli mer allvarliga än vad som tidigare predikterats. Data som

vii

presenteras i denna avhandling möjliggör modellering av Hg’s

biogeokemiska cykel på en ny detaljnivå samt möjliggör säkrare

prediktioner av hur olika ekosystem kan förväntas svara mot

miljöförändringar eller ändrad deposition av atmosfäriskt Hg.

viii

1

1. Introduction

Anthropogenic activities, such as combustion of fossil fuels and gold

mining, have significantly increased the pool of mercury (Hg) mobilized in

the environment.1-2 This has resulted in an increased chronic exposure of Hg

for marine and terrestrial organisms.3-4 Figure 1 illustrates the principal

transportation and biogeochemical processes in focus for a large community

of politicians, policy makers and scientists.5 Mercury is emitted, primary to

the atmosphere,6-8 by both natural and anthropogenic sources. A fraction of

this Hg is methylated in sediment,9 soils10 and waters11 by sulfate12-13 and/or

iron14-15 reducing bacteria to monomethylmercury (MeHg) and then

bioaccumulated and biomagnified in aquatic food webs.16-18

Consumption of fish is the main exposure rout of Hg for humans in most

regions.19 Acute poising of Hg causes severe damage to the nervous system

and in worst case death,19 such events are however nowadays rare.20 For

people with a varied diet (including fish but limited intake of certain fish

species), the exposure is typically at levels considered being without any

associated health risks. Still, 7 % of women in child-bearing age in the USA

Figure 1. Illustration of the principal biogeochemical processes of Hg, from antropogenic emission to accumulation in fish.

2

are estimated by the United States Environmental Protection Agency (US

EPA) to have a diet with Hg concentrations exceeding this level.19 MeHg is

able to cross the blood-to-brain and placenta barriers.21 Fetuses are thus

particularly vulnerable and low exposure levels of Hg have been suggested

to result in decreased birth weight and negatively affect fine motor, memory

and language skills.22 However, at present, unanimous evidence of negative

effects from low Hg dose exposure exists only for neurocognitive effects at

exposure levels typical in population groups with an unusually high intake of

fish, marine mammals or highly Hg contaminated fish from areas with local

Hg pollution sources.

In Sweden alone, tenths of thousands of lakes have Hg concentrations in

piscivorous fish exceeding the levels considered safe for human

consumption.23-24 In addition to the concerns for human and wildlife health,

Hg in fish results in financial losses (primary for commercial and sport

fisheries) and forces to changes in way of life for populations traditionally

depending on local fish sources.8,25-26 The bioaccumulated Hg primary

originates from atmospherically deposited Hg and it has been suggested, and

in some studies experimentally supported, that newly deposited Hg at a

higher extent contribute to fish Hg than past depositions do.24-25,27-28 The

atmospheric pool of Hg has increased 3-5 times since preindustrial time.29 It

has however been difficult to establish an unambiguous relationship between

recent changes in the atmospheric load and fish concentration of Hg.25,30-31 In

Sweden and in the USA a general decrease in fish Hg was observed from the

beginning of the 1980s to the end of the 1990s which was linked to

decreased atmospheric deposition rates of Hg.4,32-34 Recent studies have

however revealed a shift in this trend and increasing concentrations of Hg in

fish has been reported during the last decade.31,33-34 In Sweden, Åkerblom31

et al. reported increased concentration of Hg in pike for 36 % of lakes

(n = 25), between 1994 and 2006. The increase was positively related to total

organic carbon in the lake water. It has been suggested that variations in

hydrological conditions to (at least on a shorter time scale of decades to

centuries) possibly can counteract the positive effects expected from further

reduced atmospheric Hg load.33-34

3

Global emission inventories, although containing large uncertainties, reveals

that the anthropogenic emissions of Hg to the atmosphere have decreased

during the last decades due to implementation of emission controls and

restricted use of Hg in high income countries (mainly Europe and North

America).35

However, emissions are increasing in low income countries with

a growing economy, industrialization and an increasing demand of energy

and heat production. Pacyna et al.35-36 have predicted (for main atmospheric

emission sources of Hg in an illustrative model) that the anthropogenic

emissions could i) increase by 25% from year 2000 to 2020 if current control

practices remain unchanged and with the expected global economic growth,

ii) be reduced by 40% if currently implemented and planned emission

controls for Europe and North America are implemented worldwide; and iii)

be reduced by 55% if the technologically best available emission controls

would be implemented worldwide. United Nations Environmental Program

(UNEP) is currently leading international negotiations for a “global legally

binding instrument on Hg” (negotiations are planned to result in an

agreement open for signature in the beginning of 2013).37

The severeness of Hg as a global pollutant is to a large extent controlled by

different environmental processes resulting in MeHg formation from fresh

Hg deposits (from natural and anthropogenic emissions sources) and from

Hg stored in the environmental compartments, and the subsequent

bioaccumulation of MeHg in aquatic food webs.25,38-39 The largest fraction of

Hg emitted is accumulated and stored in deep ocean waters, or in sediment

and soils as solid and adsorbed phases of inorganic divalent Hg. Because

fresh Hg deposits and Hg stored in the environmental compartments at the

same time are contributing to the formation of MeHg formation and its

bioaccumulation, the link between current anthropogenic Hg emissions and

MeHg in fish is neither direct nor simple. It is currently uncertain to what

extent and within what timeframe a reduction of anthropogenic emissions of

Hg will result in reduced concentrations of Hg in fish. There is a consensus

among scientist, as well as environmental organizations, that an improved

understanding of Hg biogeochemistry is required to develop effective

strategies for decreasing the anthropogenic contribution of Hg in

fish.19,25,31,38,40 Unraveling the biogeochemical processes for MeHg

formation, transportation and bioaccumulation is a key objective in this

aspect. The biogeochemistry of Hg is obviously far more complex than

4

illustrated in Figure 1, involving multiple chemical forms of Hg (Table 1)

that undergo photo-, biological-, and chemical reactions and that are readily

cycled among the atmosphere, soils, wetlands, waters and sediments.38 A

chemical specie is defined by IUPAC as “specific form of an element

defined as to isotopic composition, electronic or oxidation state, and/or

complex or molecular structure” and speciation is defined as “distribution

of an element amongst defined chemical species in a system”.41 Throughout

this thesis, the sum of all possible species of inorganic divalent mercury and

monomethylmercury is referred to as HgII and MeHg, respectively. I will

refer to geochemical pools of HgII and MeHg as differing with respect to

their i) spatial distribution (e.g. buried in sediment or localized at the

sediment surface), and/or with respect to their ii) chemical speciation. In the

latter case, chemical species includes well-defined solid phases like α-HgS,

β-HgS and HgII and MeHg adsorbed to more complex components such as

natural organic matter (NOM) and iron sulphide minerals (Mackinawite,

FeS(s)).

Geochemical pools of Hg have been assumed to differ with respect to their

availability for methylation, bioaccumulation and transportation.42-45 At what

extent the different geochemical pools control Hg geo- and biogeochemistry

is however to a large extent unknown. Current insights have been derived

from experiment using quite simple model system44-46 but establishment of

the importance of the processes in complex system at conditions realistic for

natural environments are missing. This can partly be addressed to a previous

lack of suitable experimental approaches. Understanding the importance of

different geochemical HgII and MeHg pools is further complicated by the

fact that analytical methods for determination of individual Hg species in

most cases have a limit of detection (LOD) far above the concentrations

found in environmental matrices.20 We thus rely on chemical speciation

modelling,47-49 sequential extraction procedures isolating operationally

defined phases50-51 or measurement techniques requiring addition of Hg in

the lab.50,52-53 These uncertainties and lack of knowledge contributes to

MeHg and HgII often being viewed as uniform pools in sediment and soils

and the chemical speciation of solid-phase Hg is often neglected.30

Obviously, new analytical methods and experimental approaches are needed

5

Table 1. Summary of concentrations and speciation of Hg in air, water,

sediment/soils and fish (≡ denotes mineral surfaces).

% o

f H

g19

,39

,54

-55

>95

<2

<3

10

-54

30

-90

<1-3

0

<1.7

>96

0.0

6-4

<5

>95

Hg

Spe

cies

56

-58

(g

)

(O

H/C

l/B

r)n

n-2

(aq

), −

NO

M, −

(O/N

/S)≡

(OH

/Cl/

Br)

(aq

), S

H(g

),−N

OM

, −(O

/N/S

)≡

(g, a

q)

(g, a

q)

-(O

H/C

l/B

r)n

n-2

(aq

), −

DO

M(a

q),

−N

OM

, −(O

/N/S

)≡

-(O

H/C

l/B

r)(a

q),

−DO

M(a

q),

, −(O

/N/S

)≡,

(g, a

q)

(g, a

q, l

)

-(O

H/C

l/B

r)n

n-2

(aq

), (

SH) 2

(aq

), S

2H- (a

q),

S22

- (aq

), D

OM

(aq

), α

an

d β

-

HgS

(s),

-N

OM

, -(O

/N/S

)≡

−NO

M, D

OM

(aq

), S

H(g

), S

- (aq

), −

(O/N

/S)≡

(g, a

q)

−S-b

iom

ole

cule

s

−S-b

iom

ole

cule

s

Hg0

HgII

MeH

g

DM

eHg

Hg0

HgII

MeH

g

DM

eHg

Hg0

HgII

MeH

g

DM

eHg

HgII

MeH

g

[ to

tal H

g ]1

9,5

0

1-1

70

ng

m-3

0.2

-15

ng

L-1

2-2

20

0

/So

il

ng

g-1 d

.w

10

-13

00

ng

g-1 d

.w.

Air

Wat

er

Sed

imen

t

Fish

6

in order to improve our understanding of the role played by different

geochemical pools of HgII and MeHg in the biogeochemical cycle of

mercury.

This thesis involves work with a recently developed analytical method for

speciation analysis of gaseous MeHg, dimethylmercury (DMeHg), HgII and

elemental Hg (Hg0) (paper I), and a new experimental approach for

determination of Hg methylation of solid/adsorbed phase isotope tracers

(paper II-IV). Both of these methods provide great opportunities to expand

our understanding on Hg biogeochemistry and will be described and

discussed in the following chapter. Using these methods, we have studied

formation, transportation and bioaccumulation of Hg chemical forms

originating from different geochemical pools as indicated with colored

arrows (assigned paper I-IV) in Figure 2.

Figure 2. Illustration of the biogeochemical processes studied in paper I-IV (referred to by roman numerals and highlighted as colored arrows).

7

In summary, we (i) have showed a substantial emission of newly methylated

Hg from a contaminated water-sediment system (paper I), (ii) have

quantified the rates of methylation and bioaccumulation originating from

different geochemical pools of HgII (paper II-IV) in sediment slurries as well

as in larger-scale sediment-water systems with intact sediment cores, (iii)

showed that newly deposited HgII and MeHg is significantly more available

for MeHg formation and bioaccumulation than HgII and MeHg accumulated

or formed in the sediment, and (iv) showed that, in the estuarine system,

formation of MeHg from sediment pools of Hg increases with pelagic

primary production whereas enhanced loading of terrestrial NOM results in

higher formation of MeHg, as well as higher bioaccumulation, from newly

deposited Hg (paper IV).

8

2. Experimental systems and approaches

This chapter describes the possibilities and limitations of current methods

when it comes to the determination of Hg speciation in environmental

matrices. Thereafter I describe and discuss the new methods and approaches

used in papers I-IV, as well as the experimental scale (from microcosm

(papers I-II) to mesocosm (papers III-IV)).

2.1 Hg speciation and fractionation analysis in environmental

samples

Analytical methods available for quantifying concentrations of total Hg and

concentrations of specific chemical forms of Hg present at trace levels (i.e.

<100 parts per million atoms (ppma) or <100 µg g-1) in sediment, soil, water

and air (Table 1) have provided the basis from which our current

understanding of environmental Hg processes has evolved. These methods

typically include procedures for i) extraction, purification and

preconcentration of the analyte from the matrix, ii) separation of chemical

species and iii) element specific detection.20 The most commonly used

method is the determination of total Hg in solid, aqueous or gaseous

samples.50

Methods for determination of individual Hg species (e.g. the

dissolved HgII complexes Hg-(OH)2, -Cl2, and specific (-SR)2) generally

requires softer extraction procedures and/or separation techniques e.g. high

performance liquid chromatography (HPLC), which results in relatively high

LOD (typical LOD for HPLC-ICPMS of 0.0001-1 µg L-1).20 In most cases,

such as for the determination of specific HgII-thiols,56 LOD for suitable

methodologies are far above the concentrations expected in e.g. natural

water or sediment pore waters.20,56 As an alternative, fractionation, defined

by IUPAC as “process of classification of an analyte or a group of analytes

from a certain sample according to physical (e.g., size, solubility) or

chemical (e.g., bonding, reactivity) properties”, are commonly used.20,41,50

The analytical methods and modelling approaches used for Hg fractionation

and speciation in Papers I-IV are listed in Table 2.

9

Table 2. Analytical methods and model approaches used for total Hg, Hg fractionation and Hg speciation in Papers I-IV.

Matrix Analyte Method Paper Ref

Air Hg0, MeHg, DMeHg

On-line ethylation of MeHg (NaBEt4), purge and trap GC-ICPMS analysis

I 59-60

Water total Hg Sample digestion (BrCl), Hg reduction (SnCl2), purge and trap ICPMS analysis

I, II 61

total Hg Sample digestion (BrCl), on-line Hg reduction (SnCl2), ICPMS analysis

III, IV 61

HgII speciation Chemical equilibrium modeling I

MeHg Ethylation of MeHg (NaBEt4), purge and trap GC-ICPMS analysis

I-IV 62

MeHg speciation

Chemical equilibrium modeling I 47,63

Sediment total Hg Microwave assisted acid digestion (HNO3, HCl) and ICPMS analysis

I-IV 64

HgII speciation Chemical equilibrium modeling II

HgII speciation Hg LIII-edge EXAFS II 65-66

MeHg Double extraction of MeHg (CuSO4/KBr/H2SO4 followed by CH2Cl2) into MQ water, ethylation of MeHg (NaBEt4), purge and trap GC-ICPMS analysis

I-IV 67

Biota total Hg Sample digestion (80°C, HNO3 and H2SO4) ICPMS analysis

III-IV 68

MeHg Alkaline digestion (tetraethylammonium hydroxide), ethylation of MeHg (NaBEt4), purge and trap GC-ICPMS analysis

III-IV 69

Mercury in air is commonly fractionated into gaseous elemental Hg (GEM,

i.e. total gaseous Hg passing 0.45 µm filter, presumably Hg0), reactive

gaseous Hg (RGM; i.e. gaseous Hg collected in mist chambers or denuders,

presumably HgII(aq)) and particulate Hg (HgII(s), i.e. Hg collected using

filter based techniques or denuder methods).54,70-71 Elemental Hg constitutes

approximately 95 % of the atmospheric pool of Hg but RGM and particulate

HgII(s) are the primary forms of Hg deposited to land and water and hence in

many cases desirable to quantify.20

10

In water (including pore waters of sediment and soil), total Hg is often

fractionated by a separate determination of MeHg, Hg0 and HgII (as total Hg

- MeHg).61-62 The apparently dissolved fraction of MeHg and total Hg are

determined after removal of particles, most commonly larger than 0.45 µm,

by filtration.61

However, HgII and MeHg both have a strong affinity for

binding sites on solid particles and the fraction defined as “dissolved” can be

assumed to contain a significant part (12-93%)72-74 of HgII and MeHg

adsorbed to particles with a size less than 0.45 µm. Analytical methods that

enable quantification of specific MeHg and HgII species at natural

concentration levels in waters are at present not available. Chemical

equilibrium modeling approaches (thermodynamic calculations) thus play an

important role since the speciation of dissolved MeHg and HgII control the

mobility and bioavailability of Hg (paper I).47-49 As discussed by Skyllberg49,

these models still involve large uncertainties and we are far from fully

understanding Hg speciation in natural matrices. Further research is needed

to determine accurate thermodynamic stability constants and understanding

the adsorption properties of HgII and MeHg on e.g. metal sulfide surfaces.

Total Hg in the solid phase (sediment and soils) is typically fractionated by a

separate determination of MeHg and HgII.20,67,75 In samples contaminated by

mining or chloro−alkali industry, the solid phase speciation of HgII (e.g.

cinnbar, metacinnbar, HgSO4, HgO and HgCl2) can be determined using

synchrotron radiation based X-ray Absorption Spectroscopy (XAS)

techniques, e.g. X-ray Absorption Near-Edge Structure (XANES) and

Extended X-ray Absorption Fine Structure (EXAFS) Spectroscopy.76-78 For

less contaminated sediments or soils, the concentrations of Hg are however

too low and additions of Hg are needed for XAS analysis.66,79 Such addition

have however been proven useful in establishing the main adsorbed and

solid phases of HgII, as well as adsorbed phases of MeHg in sediments and

soils.49,53,66

1.1 Application of Hg stable isotope experimental approaches

For elements such as carbon and nitrogen, the occurrence of two stable

isotopes (12C, 13C and 14N, 15N, respectively) is widely used i) for

quantification purposes by isotope dilution analysis, IDA, ii) to study

reactions and processes in natural samples by addition of an isotope enriched

11

reactant or iii) to trace environmental processes in the field from natural

isotope fractionation.80-82 Mercury has as many as seven naturally occurring

stable isotopes (average natural abundance in %); 196Hg (0.14 %), 198Hg

(10.0 %), 199Hg (16.8 %), 200Hg (23.1 %), 201Hg (13.2 %), 202Hg (29.8 %) and 204

Hg (6.9 %). Multiple isotope enriched Hg tracers can thus be added to a

single system and be quantified using mass specific detection and signal

deconvolution of the detected Hg isotopic composition.25,50,60,69 This allow us

to simultaneously study several individual environmental processes in

complex experimental systems containing a “background” of ambient Hg.

Commercially available isotopically enriched Hg (often as Hg0, HgO or

HgCl2) are typically enriched in the given isotope from 30 to >99 %.50,83 In

most studies, isotopically enriched aqueous HgII or MeHg complexes are

used, however isotopically enriched DMeHg and Hg0 has also been

synthesized and used for quantification purposes.50 In this thesis also the

isotopically enriched solid phases cinnabar (α-HgS(s)) and metacinnabar (β-

HgS(s)) and, HgII reacted with Makinawite (≡FeS-HgII), and HgII and MeHg

adsorbed to natural organic matter (HgII-NOM, MeHg-NOM) have been

synthesized and used (paper II-IV).

In IDA, a known amount of an isotopically enriched Hg standard is added as

internal standard to the experimental sample before sample workup is

initiated.50 In contrast to external calibration methods, losses or incomplete

extraction of the analyte during the analytical procedures can be accounted

for (under the assumptions that added internal standard is properly

equilibrated with the sample matrix and exhibits the same (or very similar)

chemical and physical properties as the analyte). Larsson and co-workers59-

60,84 recently developed a method, applied in paper I, for determination of Hg

species/fractions in gaseous samples. By using an isotope dilution approach,

Larsson and co-workers successfully addressed several critical issues (most

importantly incomplete recoveries and species transformation) associated

with other methods.

In paper II-IV, up to 5 different isotopically enriched tracers were used to

study the reactivity and/or transportation of Hg (the 6th and 7th isotope are

reserved for detection of ambient Hg and for internal standard). The use of

isotopically enriched Hg as tracers is most commonly used to study

12

methylation and demethylation processes in sediment, soils and water.50

Typically, HgII(aq) and MeHg(aq) tracers are added to experimental samples

and incubated for up to 48 h. Thereafter the rate constants for HgII

methylation and MeHg demethylation can be calculated. The main concern

with this approach is if tracers (added as labile aqueous HgII or MeHg

complexes) will react differently than ambient Hg and MeHg, which under

relevant environmental conditions can be assumed to bind to thiol groups in

NOM, at FeS surfaces or precipitate as β-HgS(s). To address this issue we

developed an approach, discussed below, where we synthesized and added

isotopically enriched tracers as solid or adsorbed forms (paper II).

Isotope fractionation of C, N and H have been studied since the 1950’s and

is today widely used to study natural processes of these elements.85 Given

the much lower natural isotope fractionation of the Hg atoms, and a previous

lack of analytical techniques with sufficient accuracy of isotope ratio

measurements, isotope fractionation of Hg have only during the last decade

been recognized as a powerful tool to study biogeochemical processes of

Hg.86 At present a number of environmental processes have been shown to

fractionate Hg isotopes. Different anthropogenic sources of Hg are also

known to differ in isotopic signature, but to what extent isotope fractionation

could be used to e.g. track sources of Hg in the environment remains

uncertain.86-87 The approach will however most likely contribute to future

advances in our understanding of the environmental processes of Hg and

gain a wide interest among Hg scientist. Although such techniques were not

used in this thesis they are worth highlighting as emerging experimental

tools based on stable Hg isotope measurement techniques.

1.1.1 Determination of gaseous Hg0, Hg

II, MeHg and DMHg

The low concentrations of HgII, MeHg and DMeHg, in air makes

preconcentration of the analytes (typical ambient concentrations around

2-100 pg m-3 and instrumental absolute LOD about 2 pg) the main challenge

in the quantification of these chemical Hg forms.84 Available approaches,

such as liquid absorbers, cryogenic trapping and solid adsorbents suffer from

incomplete recoveries and potentially high blank levels and/or species

transformation. The use of external calibration is thus neither reliable nor

practical. By administrating isotopically enriched HgII(g), MeHg(g),

13

DMeHg(g) and Hg0(g) as internal standards to the gaseous sample, followed

by ethylation of ionic Hg species (HgII(g) and MeHg(g)) and

preconcentration on a solid sorbent (Carbotrap or Tenax), Larsson et al.,

could demonstrate reliable quantification of Hg0, MeHg, DMeHg and HgII.59-

60,84 The gaseous standards are generated from temperature controlled cells

with permeation tubes and a constant transport gas flow. The permeation

tubes contain the isotopically enriched standard as a solution or salt, and for

Hg0 as a liquid. The permeation rates will depend on temperature, the

concentration of standard, and material and dimension of the permeation

tubes. Larsson et al. could quantify species transformation during the

sampling and fairly extensive species transformations was observed.60 For

transformation during sampling of ambient air as much as 24 % of the

DMeHg(g) standard added was detected as MeHg(g) and 40 % as Hg0.

During collection of only the internal standard in inert atmosphere (N2), less

than 5 % of the DMeHg(g) standard was transformed to other chemical

forms of Hg. This emphasizes the large variances that can be expected in Hg

transformation reactions depending on the composition of the gaseous

sample.

Larsson recognized traceability as a challenge for the proposed

methodology, since no closely defined standard methods or certified

reference materials exist to assure traceability of the analytical results

gained.84 Further, the amount of internal standard added in IDA should be

matched to the expected ambient analyte concentration in the sample to

minimize the uncertainty of the quantification.88-89 Adjusting the isotope

standard additions close to the sampling occasion might however be

challenging for gaseous samples due to the time (days-weeks) needed to

obtain stable permeation rates of the gaseous standards (caused by e.g.

disturbances in adsorption-desorption equilibria of gaseous tracers on tube

walls). These challenges are however minor in comparison to the challenges

exhibited by methods not based on IDA. The method is thus a great

improvement for the determination of chemical forms of gaseous Hg and

have been successfully applied in microcosm experiments (Paper I, Björn et

al50, Larsson et al.60) and for Hg speciation analysis in natural gases (Larsson

et al.90). Further, Baya and Hintelmann91 were, for the first time, able to

quantify MeHg and DMeHg in the Arctic lower troposphere using this

methodology (average detected concentrations of 5.5 ± 2.0 and 2.3 ± 3.6

14

pg/m3 of MeHg and 2.8 ± 3.6 and 4.1 ± 2.3 pg/m3 for DMeHg in Hudson bay

and the high Arctic, respectively). They could thus support the possibility

that MeHg bioaccumulating in the Arctic system may partly originate from

MeHg and DMeHg volatilized from oceans.

1.1.2 Methylation and demethylation rates and rate constants

Methylmercury accumulated in fish originates from HgII methylated by

sulfate and/or iron reducing bacteria under anoxic conditions in soils, waters

and sediments.9-15 The concentration of MeHg in fish is thus under control of

the activity of HgII methylating bacteria, the concentration and availability of

HgII for these bacteria (i.e. speciation of dissolved, adsorbed and solid HgII),

as well as of the rate of MeHg demethylation.42 Because all these factors

vary among ecosystems, so will the expected net methylation rates of HgII.

The rates for HgII methylation and MeHg demethylation in natural

environments are traditionally described by pseudo first-order kinetic models

(Equation 1) where the rates of methylation and demethylation, rm and rd, are

calculated as the products of concentrations of HgII and MeHg, and the

methylation and demethylation rate constants, km, and kd (d-1), respectively.45

Typically the rate constants, km and kd, are experimentally determined using

isotopically enriched aqueous AHgII and MeBHg tracers. Figure 3 illustrates

the expected MeHg/HgII ratio for AHgII and MeBHg as a function of time,

following first-order kinetics (Equation 1). The rate constants are typically

determined within the timeframe when rm>>rd for AHgII and rm<<rd for

MeBHg (typically 1-48 h). By assuming negligible demethylation and

methylation of formed MeAHg and BHgII, respectively, during this time

period, km and kd can be calculated by Equation 2 and Equation 3,

respectively. For the calculation of km, the pool of HgII is assumed to be

several orders of magnitude larger than the pool of MeHg formed and can be

approximated as constant during the time of incubation.

15

Figure 3. Expected MeHg/HgII ratio as a function of time as a consequence of first-order methylation and demethylation of the AHgII and MeBHg tracers, enriched in isotope A and B, respectively. A situation is described where steady state is reached with rm = rd.

Equation 1

[ ] [ ]

[ ] [ ]

Equation 2

[MeHg]

[ ]

Equation 3

( [ ] [ ] ) [ ] )

16

In paper II, km and kd, were instead calculated by fitting a nonlinear

reversible reaction model, Equation 5 (integrated from Equation 4), to

determined MeHg concentrations (formed from a HgII tracer) as a function

of time (Figure 1 in Paper II). This approach does not require the

assumptions made in Equation 2 and Equation 3 and allows estimating rate

constants also for changes in ambient Hg. Experimental conditions may

induce a net methylation of ambient Hg as a consequence of changes in the

environmental conditions (increased temperature and mixing of sediment

slurries). In paper II, we were able to compare km determined for ambient Hg

and km determined for traditional aqueous AHgII tracers and our newly

synthesized solid/adsorbed tracers. Due to the typically small net

methylation of ambient Hg, the km values were associated with large

uncertainties and the approach is thus primary useful for method

development and evaluation purposes. In our experiment, the km values for

ambient Hg were 20-30% lower when calculated from Equation 2, as

compared when calculated from Equation 5. This suggests that by neglecting

demethylation a reasonably small (considering the experimental complexity

and typical variability associated with determination of km), however

statistically significant, error is induced. The magnitude of this error is

expected to vary considerably between samples depending on the magnitude

of the methylation and demethylation rates. The nonlinear reversible reaction

Equation 4

[

[HgII] )

[ Hg] )] [

] [

[HgII] )

[ Hg] )]

Equation 5

[ Hg] ) (

)

) [ ]

(

( ) )) [ ]

17

model requires a larger minimum number of data points compared to the

traditional approach, based on Equation 2 and Equation 3, and the latter

approach may thus be more practical in large scale surveys and monitoring

studies.

The AHgII and MeBHg tracers are typically added as labile aqueous species at

concentrations corresponding to 10-100 % of ambient MeHg and Hg.50 The

current consensus concerning the availability of HgII for uptake in bacteria

responsible for methylation is that i) dissolved complexes of HgII constitute

the available pool (neutral Hg-S and specific Hg-thiol complexes have been

suggested as particular available)44, 48, 101 and ii) precipitation and adsorption

of HgII to the solid and adsorbed phases limit the pool available for

methylation.92-94 It is generally assumed that added aqueous tracers have a

higher availability than ambient HgII.45,50 This have also been experimentally

supported (Paper II).45 The terms potential, specific or conditional rate/rate

constant have thus been used to acknowledge the presumably overestimated

km.50 For the determination of biogeochemically more relevant values on km

and kd, we have proposed solid and adsorbed phase isotopically enriched Hg

tracers to be used (Paper II, see 2.2.3)

As discussed below (see 3.2); there are reasons to consider, in contrary to the

general view, the existence of a sediment MeHg pool that is not readily

available for demethylation. Assuming the presence of HgII and MeHg pools

that are not readily available for methylation and demethylation reactions,

Equation 1 should include one second term for HgII and one for MeHg

(Equation 6) in addition to the once describing pools readily available for

methylation, HgIIA and MeHgA.

Equation 6

[[ ]

[ ]

] [ ]

[ ]

[

[ ]

[ ]

]

18

1.1.3 Determination of methylation rate constants using solid or

adsorbed chemical forms of HgII (paper II)

In paper II, we synthesized the following solid and adsorbed HgII isotope

tracers; α-199

HgS(s), β-201

HgS(s), HgII reacted with Mackinawite (≡FeS-

200HgII or ≡FeS-202HgII), and HgII bound to natural organic matter (NOM-196HgII). These tracers were incubated in brackish water estuarine sediment

slurries together with the traditionally used aqueous HgII nitrate tracer, 198Hg(NO3)2(aq). Determined average km values are shown in Equation 7 and

followed the order; β-201HgS(s) < α-199HgS(s) < ≡FeS-202HgII < NOM-196HgII

< 198Hg(NO3)2(aq). The constants were calculated using the nonlinear curve

fitting of Equation 5 and thus a km value could also be assigned for ambient

Hg (0.011 ± 0.005 d−1), calculated from a small increase in the MeHg/HgII

ratio (an experimental artifact due to disturbance of sediment layering and

increased temperature). As a comparison, a km value of 0.010 ± 0.0011 d−1

for ambient Hg was calculated from rate constants determined from

solid/adsorbed tracers and with the distribution of ambient HgII between

solid (HgS) and organically adsorbed phases (Hg-NOM) as determined

using Hg LIII-edge EXAFS. The good correspondence between the measured

and calculated values of km for ambient Hg is a strong support for the new

methodology used. It was concluded that specific solid and adsorbed HgII

tracers are much better representatives of ambient Hg than aqueous phase

tracers like 198Hg(NO3)2(aq), which yielded a km 12 times higher than the

value estimated for ambient Hg. It was furthermore shown that net

methylation of solid/adsorbed tracers, in addition to thermodynamic

properties of solid/adsorbed HgII phases, was controlled by the kinetics of

dissolution/desorption (Figure 4 in paper II).

Equation 7

[

)

)

) )]

[

[ ]

[ ]

]

19

Applying this solid/adsorbed-phase tracer methodology to other systems

requires knowledge of the chemical solid/adsorbed-phase speciation of

ambient HgII in the sample. Even though current approaches to obtain such

information (Hg-EXAFS measurements, sequential extraction methods, or

chemical equilibrium modeling) have constraints, they can be used to make

qualified assumptions about the speciation of Hg. For the sediment used in

papers II-IV, Hg LIII-edge EXAFS measurements (and speciation modeling)

suggested a solid-phase speciation of 70 % as β-HgS(s) and 30 % as Hg-

(SR)2 (paper II). In paper III and IV we demonstrate how the proposed

approach also can be applied to differentiate the reactivity of different

geochemical pools of Hg in systems of larger scale.

1.2 Scale of experiment – from micro- to mesocosms

A goal in science is to recognize and determine the causality in order to be

able to explain, predict and understand ongoing or coming natural

phenomena. This requires experiments at multiple scales and with variable

and increasing complexity.95-97 Microcosm (micro = small; cosm = world)

provides less complex and more controlled conditions, giving an opportunity

to study individual processes than cannot be singled out in larger scale

experiments. According to scaling theory, some patterns and processes will

however only be evident over a certain threshold of scale and complexity.95

The number of experiments conducted at mesocosm (meso=intermediate) or

field scale have thus increased during the last 30 years.97 The mesocosm

experiments summarized by Petersen et al.97 had a median volume of 1.7 m3

and a median duration of 49 days. As concluded by the authors, this is a

small and brief scale in relation to the physical and time scale of ecosystems.

They further identified the presence of walls as a problem since the

biological, material and energy exchange with the outside world is restricted

and the walls also provide surfaces for growth of artificial edge habitats. The

third and last main limitation of mesocosm experiment originates from

experimental artifacts arising from the decisions made by the experimental

designer (i.e. number of replicates, size, duration, control of additives etc). It

should be noted that field and whole ecosystem experiments also suffer from

scale dependence since all ecosystems vary in physical scale, manipulations

are done by experiment designer and the time scale of experiments

/monitoring is limited.

20

We studied methylation of solid and adsorbed phases of HgII in microcosm

(up to 0.005 L of sediment slurries; paper II) as well as in mesocosms

(2000 L of water and intact sediment cores; papers III-IV). Even though

absolute rates of net methylation between tracers differed, the order and the

relative differences of availability for methylation of the tracers was the

same in micro- and mesocosms (Table 3, 3.1). Similar conclusions

concerning the availability of solid/adsorbed HgII phases can thus be drawn

from both scales. This illustrates that traditional microcosm incubation for

determination of km and kd are useful experimental tools that do reflect

environmental processes occurring also in larger scale. This is not obvious

since sediments are dynamic and complex systems with a drastic change in

chemical properties with depth. Thus, mixing of sediment into slurries likely

alters the chemical and biological properties of the material. Important

differences between the microcosm sediment slurries and the mesocosm

experiment were: i) the use of intact sediment cores in the mesocosm study

(keeping the natural redox gradient and presumably bacterial community

structure), ii) establishment of a pelagic‒sediment coupling in the mesocosm

systems and maintenance of a continuous supply of autochthonous organic

carbon to the benthic zone, and iii) that atmospheric and terrestrial loadings

of HgII and MeHg could be simulated in a realistic way in the mesocosm.

The combination of molecular scale detail (i.e. solid phase speciation of HgII

and MeHg) with the more complex meso-scale experimental system help to

significantly improve our biogeochemical understanding of Hg in natural,

complex systems.

21

3. Mercury methylation, evasion and bioaccumulation

This thesis contains a summary of three experimental studies, as illustrated

in Figure 4, covering the processes of mercury methylation, demethylation

and bioaccumulation.

Figure 4. Illustration summarazing the three experimental studies in this thesis

22

3.1 Methylation of HgII

Mercury methylation in sediments is mediated by specific strains of sulfate

and iron reducing bacteria.12-15 It is generally accepted that methylation of

Hg occurs inside the cell,98-99 even if the transportation mechanisms of HgII

through the cell membrane is not clear. Different dissolved complexes of

HgII have been suggested to be available for uptake by methylating bacteria

including neutral Hg-sulfide complexes and Hg complexes with specific low

molecular mass thiols.44,48,100-101 Nano-particulate (colloidal) HgS(s) has

recently been proposed to be directly available for bacterial uptake, as

judged from studies using pure bacterial cultures.46,102 Proposed mechanisms

for uptake of such particles are however lacking. In accordance with

bacterial culture studies, we detected methylation in sediment slurries of HgII

likely originating from nanoparticles (<0.02 µm) of α-HgS(s) (paper II). Our

study showed that methylation of solid/adsorbed HgII phases is controlled by

thermodynamics and as well as kinetics of HgII dissolution/desorption. The

higher availability for methylation of nano-particualte HgS(s) than

crystalline micro-particulate HgS(s) observed in previous bacterial culture

experiments46,102, might thus also be explained by higher desorption rates

from smaller particles instead of a direct bacterial uptake of HgS(s) nano-

particles. Until further evidence for direct bacterial uptake of HgS(s)

nano-particles can be demonstrated, it is reasonable to retain the consensus

stating uptake of only dissolved HgII species.

The phase distribution of HgII in sediments is largely shifted to the solid and

adsorbed phases (typical solid-liquid partition coefficient, Kd, of Hg in

sediments is 103.5-6 (L kg-1)).43, 68 The main chemical forms of HgII in these

phases include HgII adsorbed to thiol groups in natural organic matter,65,103 or

at the surfaces of FeS(s) minerals104 and incorporated in the crystalline Hg

phases cinnbar (α-HgS) and metacinnabar (β-HgS).105 Remarkably few

studies have discussed, or studied in detail, the importance of solid phase

speciation, even though it has been generally accepted that precipitation and

adsorption of HgII are important processes limiting Hg methylation.42-45

Dissolution of HgS(s) and desorption of HgII from surfaces are to a large

extent controlled by the composition of the pore water chemistry.103,106-107

Formation rates of MeHg, derived from complex natural system at realistic

conditions, are thus required to evaluate different solid and adsorbed HgII

23

species contribution to MeHg formation. In papers II-IV we provide such

quantitative data, which previously have not been available (Table 3; net

methylation represented as the MeHg/HgII molar ratio, further discussed in

paper III). We conclude, that the solid/adsorbed phase speciation of HgII is a

major controlling factor for rates of MeHg net formation. This was shown in

traditional sediment slurries (paper II) as well as in mesocosm systems with

intact sediments (i.e. maintained natural redox gradient, presumably

maintained bacterial community structure) and a continuous supply of

autochthonous organic carbon to the benthic zone from pelagic food webs

with different structure and productivity.

The higher availability of HgII-NOM to methylating bacteria, as compared to

HgS(s) phases (papers II-IV, Table 3) can be explained by the higher

thermodynamic stability of HgS(s). As shown by EXAFS measurements,66

HgII bind to NOM by two-coordinated complexation to thiol-groups, and

when these are saturated, HgII binds to weaker O/N sites. The estimated

concentration of thiol-groups (RHS) in the NOM material used in papers II-

IV (~25 µmol RHS g-1 NOM)79 and the concentrations of HgII in the HgII-

NOM tracer (~3 µmol g-1 NOM) suggest a supply of thiol-groups large

enough for the formation of a HgII-(SR)2 complex. In the mesocosm systems

(papers III and IV), we observed higher net methylation rates in the sediment

of 204HgIIwt (204HgII(aq) added to the water of the mesocosm) than of the

Table 3. Average net methylation, as MeHg/HgII ratio (± Standard Error,SE), of HgII determined in paper II (n=5) and paper III-IV (range for 3 mesocosm treatments, n=3-9)

Paper II Papers III-IV

198HgII(aq) 0.96 (0.004) 204HgIIWt 0.026 (0.006) – 0.045 (0.01)

196HgII-NOM 0.050 (0.003) 201HgII-NOMsed 0.0077 (0.0008) – 0.0092 (0.001)

β-201HgS(s) 0.0054 (0.0007) β-200HgSsed 0.00040 (0.00003) – 0.00067 (0.0001)

α-199HgS(s) 0.0078 (0.0008)

≡FeS-200HgII 0.044 (0.004)

Ambient Hg 0.0087 (0.0002) Ambient Hg 0.0068 (0.0007) – 0.0085 (0.0008)

24

201HgII-NOMsed tracer (injected 5 mm below the sediment surface). As

discussed in paper III, we suggest that 204HgIIwt is deposited to the surface of

the sediment as primary bound to NOM particles present in the water

column and that differences in binding strength to the NOM is a contributing

cause for the higher net methylation of 204

HgII

wt compared to 201

HgII-NOMsed.

The variation in reactivity between different types of HgII-NOM complexes

is however most likely smaller (because Hg will always bind to an excess of

RSH groups with similar binding strength) than the difference in reactivity

between HgII-NOM and β-HgS(s). For HgII reacted with Mackinawite

(FeS(s)), the availability for methylation of a newly prepared tracer was

similar as of 196HgII-NOM (paper II). The availability of the ≡FeS-202HgII

tracer when it had been aged for 24 days before the incubation was however

lower and similar to the availability of β-HgS(s), suggesting a gradual

dissolution of the Mackinawite by HgII and precipitation of β-HgS(s) (as

previously shown in adsorption104 and Hg EXAFS studies79). Our results

support the idea that precipitation of β-HgS(s) in sediment is an important

“sink” for HgII that limits the bioavailability of HgII for methylating

bacteria.92-94 Pure β-HgS(s) is metastable at temperatures below 315°C but

only slowly transforms to α-HgS(s).108 The inclusion of impurities also

affects the stability of β-HgS(s) in natural systems. It is plausible that β-

HgS(s) precipitated in sediment pore water, with a complex chemical

composition, is less stable (shows a greater solubility) than commercially

available β-HgS(s) and β-HgS(s) precipitated in more pure model systems

(MQ water). Further, β-HgS(s) is (as typically for sulfur minerals) a non-

stoichiometric solid and differences in the stoichiometry (i.e. Hg/S ratio)

may be a second cause of differing availability between β-HgS(s) phases.

Indeed, differentiated solubility of β-HgS(s) phases has been demonstrated

in laboratory experiment.109

In addition to the solid/adsorbed HgII speciation, MeHg formation is

controlled by additional environmental processes limiting or enhancing the

bioavailability of HgII or the activity of methylating bacteria. In field studies,

variations in MeHg formation and concentrations among sites have been

correlated to the concentrations of dissolved sulfide, NOM and to pH.42 The

effects of these chemical factors are complex since they may affect MeHg

formation in multiple ways. For example, NOM immobilizes HgII in

sediments by adsorption42,66,106 but may also increase the solubility of β-

25

HgS(s),106 inhibit β-HgS(s) precipitation, and enhance bacterial activity,

resulting in a potentially increased methylation rate.110 Further, specific low

molecular mass HgII-thiols have been suggested to be available for

methylating bacteria.100-101 The possible HgII species accumulated and

methylated by bacteria in natural environments are still under debate mainly

due to the uncertainties in the chemical speciation of HgII in pore water.49

For sulfide, correlation to MeHg concentrations has been suggested to be an

effect of decreased availability of HgII due to a higher concentration of

charged dissolved HgII-sulfide complexes at higher concentrations of

sulfide.48,92,111 Finally, the observed correlation between MeHg concentration

in sediments and pH has been suggested to be caused by an increased

association with solid phases of HgII at lower pH values, thus limiting the

availability of HgII.112

Climate change scenarios for the Baltic Sea region predict an increased

terrestrial input of allochthonous NOM and nutrients to lakes and

estuaries.113 This may fuel the pelagic bacterial pathway114-116, reduce

phytoplankton primary production and cause a shift from a phytoplankton

based to a bacteria based pelagic food web with decreased sedimentation of

autochthonous NOM as a consequence. We hypothesized that this could

decrease the activity of methylating bacteria and thus also MeHg formation.

In the reversed direction, we hypothesized that increased sedimentation of

autochthonous NOM, caused by an increased biomass from increased

nutrients load to lakes and estuaries as a consequence of eutrophication,117

could increase the production of MeHg. It is obvious from the number of

factors influencing MeHg formation and as discussed later, its

bioaccumulation, that properly conducted meso- and large scale experiments

are needed if the effect of e.g. climate change or eutrophication is to be

predicted. In paper IV, methylation of geochemical Hg pools in mesocosm

model ecosystems with different pelagic food web structure and productivity

are discussed. In the experiment, three different treatment schemes were

applied: high and low addition of nutrients (referred to as NPhigh and NPlow,

respectively) and addition of a humic soil extract (simulating an increased

terrestrial load of humic material, referred to as HSE). For NPhigh

mesocosms, fluctuations in the net formation of MeHg from 201HgII-NOMsed

tracer correlated to that of primary production in the pelagic zone (measured

as primary production rate or chlorophyll α), whereas only small fluctuation

26

in net Hg methylation and primary production was observed in HSE

mesocosms. This shows that net methylation of HgII may respond quickly to

changes in pelagic productivity (days in these mesocosm systems). The

fluctuation in Hg methylation observed for β-200HgSsed was less pronounced

than for 201

HgII-NOMsed. This suggests that formation of MeHg from β-

200HgSsed tracer was limited by a slow rate of dissolution of HgII from the

solid phase i.e. the dissolution process was the rate limiting step.

Interestingly, net methylation of 204Hgwt (simulating recent atmospheric and

terrestrial HgII loads) increased with time in HSE mesocosm and did not

show the same pattern as the 201HgII-NOMsed tracer. The higher (however not

statistically different) average net methylation of 204Hgwt tracer in HSE was

opposite to the hypothesis proposed prior to the experiment, expecting a

lower net methylation of also recent HgII loads due to a lower primary

production. A number of different possible explanations are discussed in

paper IV. The results in paper IV again emphasize the importance of the

chemical speciation of solid/adsorbed forms of HgII. It is clear that

depending on the chemical speciation of solid/adsorbed HgII, ecosystems can

be expected to differ with respect to the outcome and magnitude of the

response to environmental changes.

3.2 Demethylation of MeHg

Processes of MeHg demethylation in surface waters include both biotic and

photo induced abiotic mechanisms.39,118-119 In sediments, demethylation has

been ascribed mainly to biological processes.39,120-121 Significant abiotic

demethylation has also been suggested, but a mechanistic understanding is

lacking.122-123 In the mesocosm systems (paper III), we report steady-state

MeHg/HgII ratios in the sediment obtained from net demethylation of the

Me198Hg-NOMsed tracer, and net methylation of HgII solid/adsorbed phase

tracers as well as of ambient Hg (during the 52 days on experiment). Based

on these quotients it was estimated that up to 30% of the Me198Hg-NOMsed

tracer was not readily available for demethylation. This observation is

consistent with previous findings from Marvin-Dipasquale et al.124 An initial

quick demethylation of added tracers (also observed in papers III-IV and by

Marvin-Dipasquale et al.), has been used in other studies to determine the

typical MeHg turnover, (km+kd)-1 of 1-3 days.125 These turnover rates are

however misguiding. They are determined from the initial decrease in MeHg

27

when the rate of demethylation is highly exceeding the rate of methylation

(rm << rd, Figure 3), as controlled by the high availability of MeHg tracer

added as an aqueous labile complex, and do not take into account the pool of

MeHg not readily available for demethylation. Instead, turnover rates

determined from the Me198

Hg-NOMsed tracer (paper III, SI) were on the

order of 90 days. It is still uncertain if the mechanism for stabilization of

MeHg (making it less available for demethylation), simply is formation of

complexes with thiol-groups in NOM, or if additional processes are

involved. As illustrated in paper III, these processes are potentially very

important for the amount of MeHg accumulated in biota and further research

to clarify mechanisms for MeHg stabilization in sediments is warranted.

3.3 Evasion of Hg

In paper I, evasion of gaseous chemical forms of Hg, i.e. Hg0, MeHg and

DMHg, was determined from sediment˗water microcosm systems containing

sediments contaminated with Hg from chloro alkali industry and pulp from

paper industry. Emission of MeHg(g) (chemical speciation modeling

suggested evasion of MeHg as a neutral CH3HgSH0 complex) was

characterized by a transient trend (Figure 2 in paper 1) with highest emission

rates reached after 12 days (110 ± 41 and 140 ± 51 fmol h-1 for ambient

MeHg and MeHg formed from added 201

HgII(aq) tracer, respectively). This

transient trend is similar to trends for MeHg concentrations in other

sediment incubation experiments (e.g. Figure 1 in paper II). Also, the

sediment concentration of ambient MeHg increased 3 times during the

course of the experiment whereas the gaseous emission of ambient MeHg

was significantly higher in the beginning of the experiment. Our data thus

suggest that gaseous emission of MeHg in sediment is controlled by recently

methylated MeHg rather than by the total pool of MeHg in sediment. Indeed,

estimated sediment−water flux of MeHg from costal marine systems reveals

variations among sites and also seasonal variations have been observed.39,126-

130 At sites in the continental shelf of southern New England, the seasonal

variation in MeHg flux was also correlated to seasonal changes in MeHg

gross production, in line with our observation.127 As previously discussed

adsorption of MeHg to e.g. NOM particles may possibly stabilize MeHg and

prevent demethylation (see 3.2). Adsorption of MeHg to solid phases may

also be the reason for the higher sediment concentration, but lower evasion,

28

of ambient MeHg compared to Me201Hg from the microcosms observed in

paper I. Treatments of the microcosms (dark, light and dark with addition of

acetate) did not cause any significant difference in MeHg evasion. For the

evasion of Hg0, the treatments suggested a predominance of photo induced

oxidation of Hg0 rather than reduction of Hg

II in this system (light radiation

from a tungsten lamp giving radiation in the visual spectral range with an

overall lower intensity than solar radiation).

The transport of Hg from sediment to water column occurs via diffusion,

advection and/or resuspension of particles.131 The importance of in situ gross

MeHg production, aqueous MeHg speciation and total concentration of

MeHg in the sediment for the sediment−water mass transfer of MeHg can

thus be expected to vary among ecosystem depending on the dominant

transportation mechanism.

3.4 Bioaccumulation of mercury

Both HgII and MeHg are accumulated in micro-sized seston but little is

known about the underlying mechanisms. Current literature suggests passive

diffusion of lipophilic mercury species such as HgCl2, CH3HgCl and low

molecular mass thiol complexes of MeHg into plankton cells.132-135

The

similar abilities of HgCl2 and CH3HgCl to pass lipophilic membranes

suggest that these chemical forms should have similar accumulation

efficiencies.134 MeHg and HgII are however distributed differently within the

cells and this result in a higher transfer efficiency of MeHg.134,136 The

fraction of total Hg occurring as MeHg thus typically increases from 0.3-

1.5% in phytoplankton to 2-22% in zooplankton (data compiled by Mason et

al.38) and >90 % in piscivorous fish.39,137

In paper III and IV, we quantified HgII and MeHg accumulated in seston and

benthic invertebrates (Chironomids, Amphipods, Polycheates and Bivalves),

after 8 weeks of experiment, from ambient Hg, and from HgII and MeHg

tracers added to the brackish water-sediment mesocosms. The accumulation

is presented as the biota−sediment accumulation factor (BSAF), which is the

29

Table 4. Accumulation of Ambient Hg (Hg occuring as MeHg (%), MeHg-BSAF and HgII-BSAF) and stable isotope analysis (δ13C and δ15N (‰)) in seston and benthic invertebrates from mesocosm experiment systems (± 1 SE).

MeHg (% of Hg)

BSAF MeHg

BSAF HgII

δ13C (‰)

δ15N (‰)

Seston 50-100 µm 8.6 (4.0) n=9 4.8 (2.2) n=9 0.39 (0.06) n=9 -23 (1.6) n=8 8.1 (0.4) n=8 100-300 µm 16 (5.3) n=8 6.7 (4.7) n=9 0.15 (0.04) n=8 -25 (1.5) n=7 8.8 (0.5) n=7 300 µm 17 (10) n=7 7.5 (4.3) n=9 0.19 (0.06) n=8 -20 (1.5) n=5 9.4 (1.1) n=5

Amphipoda 63 (11) n=3 22 (6) n=8 0.063 (0.050) n=4 -19 n=1 7.3 n=1

Polycheate 29 (9) n=9 36 (14) n=9 0.58 (0.08) n=9 -22 (0.3) n=9 9.4 (0.1) n=9

Mussel 8.6 (0.5) n=9 28 (2) n=9 2.2 (0.2) n=9 -22 (0.2) n=9 7.1 (0.1) n=9

Chironomidae 5.6 (1.5) n=4 16 (2.2) n=7 2.2 (0.3) n=5 -22 (0.3) n=5 8.5 (0.1) n=5

concentration ratio of HgII or MeHg between biota and sediment (pmol g-1

d.w./pmol g-1 d.w.). Accumulation in seston is typically given as the biota

accumulation factor (concentration ratio between biota and water), but

because the concentration in the water was below the detection limits for

most tracers, BSAF was used also for seston. The average percent of

ambient Hg occurring as MeHg, MeHg-BSAF and HgII-BSAF for ambient

Hg, and δ13C and δ15N (‰) (used to evaluated trophic and foraging

dynamics) in seston and benthic invertebrates are summarized in Table 4.

Seston (suspended material including living organism as well as inorganic

and organic particles138) was collected with plankton nets at size fractions of

50-100, 100-300 and >300 µm. The δ15N value in seston was comparable to

the values in benthic invertebrates. This is similar to findings by Eagles

Smith et al139: chironomids, amphipods and seston with a size over 80 µm

(assigned by the authors as zooplankton) had similar δ15N values whereas

the ratio in phytoplankton was lower. This suggests that primarily

zooplankton species made up the seston fractions collected (> 50 µm) from

the mesocosm. The fraction of total Hg occurring as MeHg in seston

collected (Table 4) was also within the range typical for zooplankton

(2-22 %).38

Traditional paradigms envision production of phytoplankton to constitute the

base of aquatic food webs140 and the concentrations of MeHg in fish to be a

30

consequence of dietary exposure of Hg throughout the food web.16,141

Bioaccumulation and magnification schemes and models are therefore often

based on MeHg entering the food web by accumulation in phytoplankton

from surrounding water. MeHg and HgII accumulating in pelagic plankton

will thus originate from either the sediment via advection, diffusion and

resuspention of particles or from terrestrial and atmospheric sources.131 Even

though fish concentration of MeHg typically increases with trophic level,17-18

large variations are found among ecosystems and species positioned at the

same trophic level.18,139,142 Factors regulating the production of MeHg, the

entry of MeHg into the food web as well as the trophic transfer of MeHg

have been suggested to explain these variations.139,142 The relative

importance of these factors is however poorly understood.131,142-143 Factors

suggested includes e.g. concentration of dissolved organic carbon

(DOC),110,135 pH,144 rates of primary production,145-146 atmospheric

deposition of sulfate147 and of HgII 27-28 as well as the growth rate148 and

age149 of the biota and the pelagic food web structure.18 As discussed in

paper IV, we observed higher accumulation of ambient MeHg, and MeHg

originating from Me199Hgwt and 204HgIIwt tracers in seston in HSE compared

to NPlow and NPhigh treated mesocosm. This may to some extent be explained

by a higher concentration of DOM circulating in the water column,

complexing and keeping the concentrations of Hg and MeHg added as

aqueous tracers at a higher level in the water phase of HSE mesocosms.

Even though higher rates of sedimentation were observed in HSE

mesocosm, the sedimentation of Hg and MeHg obviously was overridden by

the effect of higher concentrations of DOC in the water column per se. As

discussed in paper IV, this was likely not the only explanation for the higher

MeHg concentrations in seston originating from Me199Hgwt and 204HgIIwt

tracer. Additional plausible factors that may have contributed include algal

bloom dilution150, a higher degree of accumulation of MeHg-DOM

complexes by bacteria and/or larger biomagnification due to an additional

trophic level151 introduced in HSE mesocosm.

Several studies have stressed the importance of pelagic food web couplings

to the benthic zone for trophic transfer of nutrients and energy152 as well as

MeHg.137,139,142-143 Benthic organisms can accumulate MeHg from the

dissolved MeHg fraction and via sediment ingestion.153 Aquatic organisms

feeding on benthic invertebrates, so called benthivores, thus provide a

31

secondary pathway for the transfer of MeHg from sediment to the pelagic

food web (in addition to MeHg entering the food web via accumulation in

phytoplankton after transfer of MeHg from sediment to water). In the

reversed direction, benthic organisms feeding on deposited planktonic

material and necrophages may forage on fish carcasses with a high MeHg

concentration. Together, the pelagic-to-benthic and the benthic-to-pelagic

foraging results in a trophic feedback cycle of MeHg that not only preserves

MeHg within the biota compartment, but may also result in higher MeHg

levels in benthivores and piscivorous fish.142 Figure 5 illustrates the trophic

transfer pathways in the aquatic food web.

Amphipods feed both on suspended particulate matter in the water column

by filtration, and detritus and sediment by deposit feeding.154-155 Amphipods

showed higher MeHg-BSAF and HgII-BSAF for Me199Hgwt tracer, and lower

for Me198Hg-NOMsed added 0.5 cm below the sediment surface, in

comparison to the other benthic invertebrates (Figure 6 and Figure 7).

Figure 5. Illustration of biaccumualtion and biomagnification of MeHg from sediment and water to the benthic and pelagic foodweb. Solid lines represents abiotc transportation processes and dashed lines bioaccumulation and magnification.

32

This suggests that the Amphipods to a larger extent feed on material

originating from the water column in comparison to the other benthic

invertebrates. Chironomids and Bivalves (mussels), well-known filter

feeders, had similar concentrations of MeHg and HgII originated from the

tracers and from ambient Hg. They also had a similar fraction of Hg

occurring as MeHg (Table 4). Polychaetes (ragworms) had higher %C, %N,

δ15N (‰) as well as a generally higher MeHg-BSAF of sediment tracers.

This suggests that Polychaetes, by their digestion of sediment, were exposed

for higher concentrations of MeHg than chironomids and bivalves that

primary filter feeds. Sediment digestion as a primary exposure route of Hg

for Polychaetes have also previously been shown.153

As shown in Figure 6 and Figure 7, HgII and MeHg from tracers added to the

water (Me199Hgwt and 204HgIIwt) accumulated to a higher extent than tracers

injected into the sediment (MeHg196Hg-NOMsed, 201HgII-NOMsed and β-

200HgSsed) in both seston and benthic invertebrates. We therefore conclude

newly imported HgII and MeHg from terrestrial and atmospheric sources to

be more available for accumulation in seston and benthic invertebrates than

HgII and MeHg previously accumulated in the sediment or MeHg formed in

situ from previously accumulated HgII. Furthermore, benthic invertebrates

accumulated MeHg from the MeHg196Hg-NOMsed tracer and MeHg formed

from 201HgII-NOMsed and β-200HgSsed tracers to a much higher extent than

seston. This suggests that benthic invertebrates may play an important role in

transferring MeHg accumulated or in situ formed in the sediment to the

pelagic food web. Our results in papers III and IV thus emphasize that the

traditional bioaccumulation and biomagnification pathways of mercury

(MeHg in water → phytoplankton → zooplankton → fish) are too

simplified.

33

Figure 6. MeHg and HgII accumulation from tracers added 0.5 cm below

the sediment surface in seston and benthic invertebrates (shown as MeHg-

and HgII-BSAF ± 1 SD)

34

Figure 7 MeHg and HgII accumulation from tracers added to the water

column in seston and benthic invertebrates (shown as MeHg- and HgII-

BSAF ± 1 SD)

35

4. Implication and final remarks

In papers I-IV we concluded that:

the solid and adsorbed speciation of HgII is a principle factor

controlling the rate and net methylation of Hg terrestrial and atmospheric sources of HgII and MeHg are more

available for methylation, as well as for bioaccumulation, than

sediment pools of HgII and MeHg a fraction of MeHg in sediments is not readily available for

demethylation or evasion, but still available for bioaccumulation in

benthic invertebrates

Extensive efforts have been undertaken to reduce anthropogenic emissions

of Hg since the health effects of MeHg showed in e.g. pollution catastrophes

(Minamata, Japan, 1950’s and Iraq, 1970’s).35,156 These actions have indeed

resulted in decreased atmospheric depositions of Hg and downward trends of

mercury concentration in fish in Sweden and the USA, for examples, have

also been observed.4,32-34,157 While further efforts are planned, negotiated and

under implementation there are still large uncertainties how further

decreased anthropogenic Hg emissions will affect MeHg concentrations in

fish.37 Two main concerns are that i) the response time for levels of MeHg in

fish may be on the order of decades or up to centuries due to a continuous

leakage and export of Hg stored in soils, and that ii) decreased

anthropogenic depositions may be counteracted by other environmental

changes such as climate changes and eutrophication.25,31,33

In the METALLICUS (Mercury Experiment to Assess Atmospheric Loading

in Canada and the United States) project, the response time of reduced

atmospheric depositions has been simulated in mesocosm and whole

ecosystem scale experiments.25,27-28 It was concluded that HgII deposited

directly to the water surface was quickly methylated. The increase of the

concentrations of Hg in fish was immediate, suggesting a very fast response

time between changes in rates of atmospheric Hg deposition and Hg

concentrations in fish. As pointed out in the METALLICUS studies, the fast

response time is only true for systems receiving the major Hg input from the

36

atmosphere as a load directly onto the water body, and not indirectly via

runoff from terrestrial ecosystems. For aquatic systems with a relative large

watershed area, the response time may be delayed up to centuries due to the

legacy of Hg stored and gradually leached and exported from soil. In the

METALLICUS whole ecosystem study, isotopically enriched Hg were also

added to the watershed (forest and wetland), however these loadings could

not, during 3 years after first load addition, be detected in the biota.25 Table 5

summarizes literature data on the relative contribution from terrestrial and

atmospheric sources of Hg (note: also including Hg0) and MeHg to different

aquatic systems. As shown, terrestrial imports of Hg is higher than

atmospheric imports in most estuaries and lakes where the loadings of Hg

has been modeled.

Transfer of MeHg from sediment to the water systems, determined using

flux chambers or pore water MeHg concentration gradients, is typically

interpreted as net methylation of Hg accumulated in sediments.39, 129-130

Thus, the possible mobilization of MeHg from sediment into the water phase

is neglected. We show that HgII recently deposited to the sediment surface is

more readily available for methylation processes than the already

accumulated pool of HgII (paper III). It is also shown that newly formed

MeHg was the main source of MeHg evasion from the sediment system

(paper I). Based on our results, we suggest that recently (months-years)

deposited HgII and MeHg contribute to a significantly higher degree to

sediment-water fluxes than HgII and MeHg already stored in sediments for a

longer time. Separating the contribution from recent and past Hg deposited

deposition from field observations is however challenging. It emphasizes the

need for quantitative data on the availability of specific geochemical pools

of Hg for methylation and bioaccumulation. Such data were successfully

derived from our mesocosm scale system, allowing modeling of MeHg

formation and bioaccumulation in unprecedented detail. Studies in the

METALLICUS project are loading experiments where the atmospheric

deposition of Hg is simulated by isotopically enriched, aqueous HgII tracers.

Thus, the study is restricted to a comparison of the reactivity of recent, labile

Hg loadings with the reactivity of the entire Hg pool (i.e. ambient Hg)

without differentiating between younger and older fractions (corresponding

to recent and past loadings) of the sediment pool.

37

Table 5. Terretrial/atmospheric ratios for import of total Hg and MeHg to the water body of oceans, estuaries and lakes. The total contribution from terrestrial and atmospheric sources to the total Hg and MeHg import to the water are given in paranthesis. Terrestrial/Atmospheric input ratios

(100·(Terrestrial+Atmospheric)/(Total inputs)) Ref total Hg MeHg

Ocean Global (pre. industrial) 158 0.030 (100 %) Global (current) 158 0.065 (100 %) Estuaries Global Estuarine 159 1 (98 %) − San Francisco Bay 160 1.5 (6.4 %) − (<4 %) Tokyo Bay 161 1.9 (19 %) − Long Island Sound 162 7.9 (100 %) 6.9 (33%) NY/NJ Harbor Estuary 163 31 (100 %) 27 (76 %) Chesapeake bay 164 1.6 (79 %) 4.2 (30 %) Lakes Superior 165 0.6 (87 %) only terr. (26 %) Michigan 165 0.31 (44 %) − (0 %) Onondaga 165 310 (91 %) only terr. (6 %) Little Rock Lake 165 only atm. (40 %) − (0 %) Clear Lake 166 188 (100 %) − Lake Champlain 167 1.6 (100 %) − Big Dam West 168 10 (74 %) only terr. (91 %) Spring Lake 169 0.09 (93.9 %) 0.51 (15 %)

In papers III-IV, we constructed simple mass balance budgets to illustrate

different scenarios in an estuarine environment. (Figure 8, Paper III (Figure

5 and Table 1), Paper IV (Figure 5)). The origin of MeHg in sediment and

biota was calculated from the determined ratios of net methylation and

demethylation, and the Biota Sediment Accumulation Factors (BSAF)

derived from specific HgII and MeHg tracers. These tracers simulated more

recently imported HgII and MeHg from atmospheric and terrestrial sources,

as well as older, stored sediment pools of HgII and MeHg. A challenge when

applying these types of models is the selection of scales, including the time

and the depth in the sediment at which methylation and demethylation

reactions are considered most active. We used calculated cumulative

terrestrial and atmospheric loads of HgII and MeHg, typical for estuarine

environments during a 2 months period, and an active sediment layer of 1.5

cm in our model to match the experimental scale from which our

quantitative data was derived. Since sediments store the largest reservoirs of

38

Hg in lakes and estuaries, the mass of sediment considered as “active” can

be expected to largely impact the results of Hg mass balance budgets. For

estuarine mass balance budgets, “active” sediment layer of 1 and 15 cm has

previously been assumed for Chesapeake and San Francisco Bay,

respectively160,164

and the timescale is typically year-1

.160-164

It is easy to understand that it has been difficult to establish a clear

relationship between MeHg in fish and the concentrations of Hg or MeHg in

sediment among ecosystems when comparing the mass budget of MeHg in

sediment and biota (Figure 8) and by looking at the variations of terrestrial

and atmospheric inputs among ecosystem (Table 5). Because of differences

in the availability for methylation and bioaccumulation, specific

geochemical Hg pools give very different predicted contributions to MeHg

in sediment, seston and benthic organisms (Figure 8, Paper III (Figure 4-5)).

Since such data have not previously been available, we need to reevaluate

the importance of sediment pools versus newly imported Hg from

atmospheric and terrestrial sources. Most urgently, there is an obvious risk

Figure 8. Mass balance budget calculations (denoted model A, in paper III) for pools of MeHg in sediment and biota (seston, Amphipodas and benthic invertebrates) consiting of newly imported MeHg and HgII from terrestrial and atmospheric sources and older sediment pools of MeHg and HgII (having a composition of 70 % β-HgS and 30 % Hg-NOM).

39

that contributions of MeHg in fish from terrestrial sources of MeHg and HgII

are underestimated. From the different Hg loading and solid phase speciation

scenarios postulated (Paper III) we conclude that at constant loading rates, it

can be expected that the sediment pools’ contribution to MeHg in sediment

and biota is governed by the solid-phase speciation of HgII and the formation

of MeHg pools not readily available for demethylation. In paper IV, we

calculated mass balance budgets for MeHg in sediment and seston under the

scenarios simulated in NPlow, NPhigh and HSE mesocosm. Although

associated with large uncertainties, the model suggested that at predicted

climate change scenarios for Scandinavia, the concentration of MeHg in

plankton could double due to increased inflow of MeHg and HgII and an

increased availability of Hg for MeHg formation and bioaccumulation. For

the eutrophication scenario, only small changes in sediment and seston

concentrations of MeHg was predicted.

Even though the availability of geochemical pools of HgII and MeHg for

methylation, demethylation and bioaccumulation will differ among sites, our

data can, until similar data have been obtained also from other ecosystems,

be used to improve our understanding of Hg biogeochemistry and to predict

the effect of environmental changes or changes in Hg loading rates. As

future outlooks following this thesis work I suggest some key research

questions to address important knowledge gaps in Hg biogeochemistry

related to environmental changes, atmospheric and terrestrial loading rate

changes as well as Hg and MeHg bioaccumulation in fish:

Determination of methylation rates of solid/adsorbed phase tracers

and solid phase speciation in other types of ecosystems. Such data

could allow an evaluation to what extent differences in the solid

phase speciation of Hg among ecosystems could explain differences

in sediment as well as biota concentrations of MeHg.

Studying mechanisms and processes leading to accumulation of

persistent MeHg pools in sediments. Such knowledge is at the

present scant and the importance of such pools in sediments is not

well recognized.

Application of the mesocosm scale experiment on fresh water

system. We experimentally demonstrated how environmental

40

changes such as global warming and eutrophication may affect

MeHg formation and bioaccumulation of Hg in an estuarine system.

Such data is also needed for freshwater ecosystems where the

responses might differ in direction and magnitude.

Addressing these three key questions will broaden our understanding of

Hg biogeochemistry and help understand ways to mitigate the

anthropogenic contribution of MeHg in the environment.

41

Acknowledgements

Mathias, jag vill tacka dig mest av allt. För allt du gav och du lärde mig, jag

vet att du hade varit stolt. Vi hade säkert suttit o diskuterat små detaljer i

avhandlingen utan att förstå varandra alls (o till slut kanske insett att vi

menade samma sak). Tack för att du var min allra bästa vän och livskamrat

under så många år, tack för vår underbara dotter. Jag kommer alltid älska dig

och berätta för din dotter om vilken omtänksam och underbar pappa hon

haft. Vi saknar dig väldigt här nere!

Tack Ida, du är underbar! De finns så många stunder som jag inte skulle

klarat av utan dig! Tack för all stöttning och att du låtit mig vara en sån stor

del av ditt liv.

Ett stort tack till min o Frejas familj som stöttat och hjälpt med allt från

kärlek o värme till hjälp med barnpassning o alla små ting: mamma, pappa,

Eva, svärmor Susanne, svärfar Eskil, bror Jens med familj, Mathias

syskon med respektive och övrig familj till Mathias, moster Kerstin.

I would like to thank the staff at the Department of Chemistry and Umeå

Marine Sciences Centre for providing me extra support and help so I could

come back to work again and finish my thesis. I felt a great support and a

genuine care from you. A special thanks to my supervisors and co-authors

(especially Erik B, Ulf, and Mats) for listening and for the extra support.

During my time in Umeå, I always felt I was surrounded by people who

were true friends, and people that would support me if I ever needed it.

When my life then was turned upside down, I got at so much more support

than I ever could have imagined. Also from friends I just got to know or that

I lost contact with. So I would like to thank all of you who in small and big

ways showed your support. Even if I did not have time to have all the

coffees, dinners and movie nights you suggested, the gesture meant a lot to

me.

42

I would like to thank the Department of Chemistry and Umeå Marine

Sciences Centre for giving me the chance to opportunity to do research.

Tack till min underbara huvudhandledare Erik! Tack för att du gav mig

vägledning såväl utrymme att utvecklas och växa. Tack för allt roligt vi haft

tillsammans under de här åren. Min tid som doktorand har varit spännande,

rolig och helt fantastisk, och mycket av det beror på att du lagt ner din tid

och själ i våra forskningsprojekt och i att vägleda mig i forskningen. Tack

också för att du varit en av mina närmsta vänner och stöttat mig i både

framgångar och motgångar. Tack för skratten, guidningen in i vinernas värld

och inte minst alla fartfyllda och legendariska stunder på dansgolvet!

Ulf, tack för att du delar med dig så mycket av din passion för forskning. Det

är verkligen inspirerande och jag är väldigt glad att jag fått jobba med dig.

Tack också för att du, trots ett fullspäckat schema, tog dig tiden att läsa min

avhandling och ge feedback. Erik L, tack för support och all praktiskt hjälp

med galna mesocosm-idéer. Någon uppskjutning av en mesocosm till

rymden blev det inte (och inte heller sediment-tårta), men de kändes som vi

fixade saker som tekniskt var i samma klass. Tack till mina medförfattare:

speciellt Mats för ditt engagemang och stöttning och Agneta, utan din

kompetens hade mesocosm-projektet inte varit möjligt. Tack Tom, för att du

guidade mig under första tiden som doktorand, det blev ett stort tomrum i

gruppen sen du slutade och jag saknar din humor och personlighet. Thanks

to my super-mesocosm-team including Minh and Helen!!! The mesocosm

project and my time as a PhD student would not have been the same without

the two of you!!! Tack Lasse för två oförglömliga sommar-projekt i

Vietnam och Kambodja. Tack till övriga kollegor i forskningsgruppen:

Lars, Sylvain, Andreas, Max, Rose-Marie, Solomon och Wolfgang. Tack

Yvonne för tiden som kontorskompisar och för din vänskap.

Tack till alla er andra som bidragit på olika sätt till arbetet som ligger till

grund för artiklarna: Tom Larsson, Anna Karlsson, Ida Tjerngren, Johan

Nordbäck (Swedish Geotechnical Institute), Dan Boström, Per Hörstedt,

Per-Olof Westlund, Andrey Shchukarev, Staffan Åkerblom och framför

allt ett stort tack till Henrik Larsson och övrig teknisk personal på UMF.

43

Har så många underbara vänner att tacka och jag kommer inte (dagen innan

tryck) kunna komma ihåg att ta med alla. Tack till Åsa o Frasse, Sandra

och Erik, Tuva-Li och Nicklas för barnpassning, trevligt sällskap och goa

middagar/luncher. Tack Åsa, också för ditt stöd sista halvåret. Tack till alla

som blivit, eller varit, del av ”analytisk kemi” för trevliga fikaraster och

vilda analyt-fester! Tack Anna N för att du visade mig Australien. Tack

Lisa och Thomas för trevliga spelkväller och goda middagar. Tack

Andreas för din vänskap och att du berikat både mitt och Mathias liv. När

jag tänker på dig som vän tänker jag på när du, flera kvällar i rad, kom till

mig på sjukhuset och satt o skrattade med mig framför TVn, du gjorde

verkligen skillnad då! Tack Sarah för sällskap till magdansen, hoppas vi får

chans att dansa ihop fler gånger. Tack Kristina, Jonas, Emil, Josefina,

matgruppen och alla andra vänner som berikar mitt liv!

Till sist, Tack till min underbara och älskade dotter Freja! Du och jag har

en underbar liten familj ihop!

44

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