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 (αHgS(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
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
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
Mercury methylation, evasion and bioaccumulation
21
3.
4.
8
3.1
Methylation of HgII
22
3.2
Demethylation of MeHg
26
3.3
Evasion of Hg
27
3.4
Bioaccumulation of mercury
28
Implication and final remarks
35
Acknowledgements
41
References
44
i
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.
ii
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 indepth modeling of processes in the Hg biogeochemical cycle that will
iii
improve our understanding and the predicting power on how aquatic
ecosystems may respond to environmental changes or differences in loading
rates for atmospheric Hg.
iv
Abbreviations
BAF
BSAF
DMeHg
DOM
d.w.
EXAFS
GC
GEM
HPLC
HgII
ICPMS
IDA
IUPAC
KD
kd
km
LOD
MeHg
METALLICUS
NOM
rd
RGM
rm
Sed
UNEP
US EPA
Wt
XAS
XANES
Biota Accumulation Factor
Biota-Sediment Accumulation Factor
Dimethylmercury
Dissolved Organic Matter
Dry Weight
Extended X-ray Absorption Fine Structure
Gas Chromatography
Gaseous Elemental Mercury
High Performance Liquid Chromatography
Inorganic divalent mercury
Inductively Coupled Plasma Mass Spectrometry
Isotope Dilution Analysis
International Union of Pure and Applied Chemistry
Solid-liquid partition coefficient
Demethylation rate constant
Methylation rate constant
Limit of Detection
Monomethylmercury
Mercury Experiment to Assess Atmospheric Loading in
Canada and the United States
Natural Organic Matter
Rate of demethylation
Reactive Gaseous Mercury
Rate of methylation
Sediment
United Nations Environmental Program
United States Environmental Protection Agency
Water
X-ray Absorption Spectroscopy
X-ray Absorption near-edge Structure
v
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
vi
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.
vii
viii
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.
1
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
2
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
3
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
4
5
(g, aq)
-(OH/Cl/Br)nn-2(aq), −DOM(aq), −NOM, −(O/N/S)≡
-(OH/Cl/Br)(aq),−DOM(aq),, −(O/N/S)≡,
(g, aq)
Hg0
HgII
MeHg
DMeHg
0.2-15
ng L-1
DMeHg
Hg
MeHg
10-1300
ng g-1 d.w.
Fish
−S-biomolecules
>95
<5
(g, aq)
MeHg
−S-biomolecules
0.06-4
−NOM, DOM(aq), SH(g), S- (aq), −(O/N/S)≡
HgII
ng g-1 d.w
/Soil
II
>96
-(OH/Cl/Br)nn-2(aq), (SH)2(aq), S2H-(aq), S22-(aq), DOM(aq), α and βHgS(s), -NOM, -(O/N/S)≡
Hg
30-90
10-54
<3
2-2200
(g, aq, l)
<1.7
(g, aq)
DMeHg
0
<1-30
(OH/Cl/Br)(aq), SH(g),−NOM, −(O/N/S)≡
MeHg
<2
(OH/Cl/Br)nn-2(aq), −NOM, −(O/N/S)≡
HgII
>95
% of Hg
ng m-3
(g)
56-58
Hg0
Hg Species
1-170
19,50
Sediment
Water
Air
[ total Hg ]
19,39,54-55
Table 1. Summary of concentrations and speciation of Hg in air, water,
sediment/soils and fish (≡ denotes mineral surfaces).
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 IIV (referred to by roman numerals and highlighted as colored arrows).
6
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).
7
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.
8
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
Hg , 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
total Hg
Microwave assisted acid digestion (HNO3,
HCl) and ICPMS analysis
I-IV
64
HgII speciation
Chemical equilibrium modeling
II
Hg 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
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
Sediment
0
II
Biota
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
9
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
10
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 (βHgS(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-workers5960,84
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
11
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, HgII, 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),
12
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.5960,84
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
13
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.
14
Figure 3. Expected MeHg/HgII ratio as a function of time as a consequence of firstorder 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
( [
]
[
15
])
[
]
)
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] )
)
(
(
)[
)
(
16
)) [
]
]
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
[
[
[
]
]
]
[
[
17
]
]
[
[
[
]
]
]
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-199HgS(s), β-201HgS(s), HgII reacted with Mackinawite (≡FeS200
HgII or ≡FeS-202HgII), and HgII bound to natural organic matter (NOM196
HgII). These tracers were incubated in brackish water estuarine sediment
slurries together with the traditionally used aqueous HgII nitrate tracer,
198
Hg(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
)
)
[
[
[
)
)]
18
[
]
]
]
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.
19
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 Hg II
and MeHg) with the more complex meso-scale experimental system help to
significantly improve our biogeochemical understanding of Hg in natural,
complex systems.
20
3. Mercury methylation,
bioaccumulation
evasion
and
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
21
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) nanoparticles. 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
22
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 IIIV (~25 µmol RHS g-1 NOM)79 and the concentrations of HgII in the HgIINOM 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
198
196
β-
Hg (aq)
II
Hg -NOM
201
α-
Papers III-IV
II
HgS(s)
199
HgS(s)
≡FeS-
200
Hg
II
Ambient Hg
0.96 (0.004)
204
HgIIWt
0.050 (0.003)
201
II
0.0054 (0.0007)
β-
0.026 (0.006) – 0.045 (0.01)
Hg -NOMsed
0.0077 (0.0008) – 0.0092 (0.001)
200
0.00040 (0.00003) – 0.00067 (0.0001)
HgSsed
0.0078 (0.0008)
0.044 (0.004)
0.0087 (0.0002)
Ambient Hg
23
0.0068 (0.0007) – 0.0085 (0.0008)
201
HgII-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 204HgIIwt compared to 201HgII-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 βHgS(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 nonstoichiometric 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 β24
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
25
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 201HgII-NOMsed. This suggests that formation of MeHg from β200
HgSsed 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
26
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 Me198Hg-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 201HgII(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.39,126130
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,
27
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 HgII 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 0.31.5% 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
28
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
100-300 µm
300 µm
8.6 (4.0) n=9 4.8 (2.2) n=9 0.39 (0.06) n=9
n=8
n=9
n=8
16 (5.3)
6.7 (4.7)
0.15 (0.04)
n=7
n=9
17 (10)
7.5 (4.3)
0.19 (0.06) n=8
-23 (1.6) n=8
n=7
-25 (1.5)
-20 (1.5) n=5
8.1 (0.4) n=8
n=7
8.8 (0.5)
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
29
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
30
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.
31
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 β200
HgSsed) 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.
32
Figure 6. MeHg and HgII accumulation from tracers added 0.5 cm below
the sediment surface in seston and benthic invertebrates (shown as MeHgand HgII-BSAF ± 1 SD)
33
Figure 7 MeHg and HgII accumulation from tracers added to the water
column in seston and benthic invertebrates (shown as MeHg- and HgIIBSAF ± 1 SD)
34
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
35
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.
36
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)
Global (current)
Estuaries
Global Estuarine
San Francisco Bay
Tokyo Bay
Long Island Sound
NY/NJ Harbor Estuary
Chesapeake bay
Lakes
Superior
Michigan
Onondaga
Little Rock Lake
Clear Lake
Lake Champlain
Big Dam West
Spring Lake
158
0.030 (100 %)
0.065 (100 %)
158
159
160
161
162
163
164
165
165
165
165
166
167
168
169
1 (98 %)
1.5 (6.4 %)
1.9 (19 %)
7.9 (100 %)
31 (100 %)
1.6 (79 %)
−
− (<4 %)
−
6.9 (33%)
27 (76 %)
4.2 (30 %)
0.6 (87 %)
0.31 (44 %)
310 (91 %)
only atm. (40 %)
188 (100 %)
1.6 (100 %)
10 (74 %)
0.09 (93.9 %)
only terr. (26 %)
− (0 %)
only terr. (6 %)
− (0 %)
−
−
only terr. (91 %)
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
37
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).
38
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 Hg II 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
39
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.
40
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.
41
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.
42
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!
43
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