Determination of mercury chemical
speciation in the presence of low
molecular
mass
thiols
and
its
importance for mercury methylation
Van Liem-Nguyen
Department of Chemistry
Doctoral thesis
Umeå, 2016
i
© Van Liem-Nguyen, Umeå 2016
This work is protected by the Swedish Copyright Legislation (Act 1960:729)
ISBN: 978-91-7601-469-1
Cover: boreal wetland by Van Liem-Nguyen
Electronic version available at http://umu.diva-portal.org/
Printed by: The service centre in KBC, Umeå University, Sweden 2016
ii
To my family
To my beloved wife
iii
iv
Abstract
Methylmercury (MeHg) is a neurotoxic compound that threatens the well-being of
humans and wildlife. It is formed through the methylation of inorganic mercury (Hg II)
under suboxic/anoxic conditions in soils, sediment and waters. The chemical speciation of
HgII, including specific HgII species in aqueous and solid/adsorbed phases, plays a key role
in MeHg formation. Chemical forms of HgII which have been reported to be available for
uptake in methylating bacteria include neutral HgII–sulfide complexes, HgII complexes
with specific low molecular mass (LMM) thiols, and nanoparticulate HgS(s). Accurate
determination of the chemical speciation of HgII is thus crucial when elucidating the
mechanism of MeHg formation. The concentration of HgII–LMM thiols complexes is
predicted to be extremely low (sub fM range). Current analytical methods do not allow
direct quantification of HgII complexes due to the very low concentration of these
complexes, and therefore determination rely on thermodynamic modeling. Accurate
stability constants for HgII–LMM thiols complexes and quantification of LMM thiol
ligands in environments are thus required to precisely determine the concentration of
such complexes.
In this thesis, a novel analytical method was developed based on online pre-concentration
coupled with liquid chromatography tandem mass spectrometry to determine the
concentration of 16 LMM thiols (Paper I). This method was successful in detecting 8 LMM
thiols in boreal wetland porewaters, with mercaptoacetic acid and cysteine being the most
abundant. The total concentration of individual detected LMM thiols ranged from sub nM
(LOD=0.1 nM) to 77 nM. Moreover, the stability constant (β2) for HgII complexes with 15
LMM thiols were directly determined for the first time by competing ligand exchange
experiments combined with liquid chromatography ICPMS analysis (Paper II). Values of
log β2 for the reaction Hg2+ + 2LMM-RS- = Hg(LMM-RS)2 ranged from 34.6 for. Based on
the determined constants of Hg(LMM-RS)2 complexes and state-of-the-art constants from
literature for other HgII complexes, we established comprehensive thermodynamic
speciation models for MeHg and HgII in boreal wetlands (Paper III). The speciation of HgII
was coupled with the HgII methylation rate constant (km) determined with different
enriched Hg isotope tracers (Paper IV). There was a good correlation (R 2=0.88) between
the km determined by a HgII(aq) tracer added as Hg(NO3)2 with high bioavailability and a
tracer where HgII was bond to thiol groups in natural organic matter (HgII-NOM(ads)) and
has a lower bioavailability. The HgII(aq) tracer was consistently methylated at 5 times
higher rate than the HgII-NOM(ads) tracer. A good correlation was observed between the
concentration of biologically produced LMM thiols and km in the boreal wetlands. In a
mesocosm study of estuarine sediment-brackish water systems, increased concentration of
phytoplankton chlorophyll α due to macro nutrient additions led to an increase in HgII
methylation rate of the HgII(aq) but not of the HgII-NOM(ads) tracer or ambient HgII
species (Paper V). Furthermore, simulated newly deposited HgII species from atmospheric
and terrestrial sources were exhibited significantly higher HgII methylation rates when
compared with simulated aged sediment HgII pools. Through the development and
adoption of novel analytical methods, this thesis reveals the significance of LMM thiols in
Hg biogeochemistry by precise determination of HgII–LMM thiol complexes in natural
environmental systems.
v
vi
Table of Contents
List of publications ................................................................................................ ix
Abbreviation........................................................................................................... xi
1.
Introduction .........................................................................................................1
1.1. Mercury ..............................................................................................................1
1.2. Chemical species and chemical speciation of mercury .................................. 2
1.3. Methylation of HgII .......................................................................................... 3
1.4. Aims of the thesis ............................................................................................. 5
2.
Materials and experimental approaches ........................................................... 6
2.1. Study sites......................................................................................................... 6
2.1.1. Boreal wetlands ......................................................................................... 6
2.1.2. Coastal marine system .............................................................................. 7
2.2. Experimental approaches ............................................................................... 7
2.2.1. Determination of LMM thiols in natural waters ..................................... 7
2.2.2. Determination of total thiols in natural waters and soils....................... 8
2.2.3. Determination of total Hg and MeHg in natural waters and soils ........ 9
2.2.4. Determination of the methylation and demethylation rate constants .. 9
2.2.5. Determination of stability constant for Hg(LMM-RS)2 complexes ..... 10
2.2.6. Chemical speciation of MeHg and HgII in boreal wetlands ................. 10
2.2.7. Mesocosm experiment............................................................................. 11
2.2.8. Mercury isotope tracers and nutrient preparation ................................ 11
3.
Results and Discussion ..................................................................................... 13
3.1. Determination of LMM thiols in boreal wetland porewaters (Paper I) ...... 14
3.1.1. Limits of detection of the method ...........................................................15
3.1.2. Determination of LMM thiols in boreal wetland porewaters ................15
3.2. Determination of stability constants for Hg(LMM-RS)2 complexes (Paper
II) ........................................................................................................................... 16
3.3. Mercury chemical speciation in the boreal wetlands (Paper III) ............... 18
3.3.1. Geochemical properties at the wetland sites ......................................... 19
3.3.2. Chemical speciation model for MeHg ................................................... 19
3.3.3. Chemical speciation model for HgII ....................................................... 21
3.3.4. Formation of HgS(s) ............................................................................... 23
vii
3.4. Factors controlling MeHg formation (Paper IV &V) ................................... 25
3.4.1. Determination of HgII methylation rate constant ................................. 25
3.4.2. Effect of microbial activity on labile HgII tracer added ex situ ............ 26
3.4.3. Effect of microbial activity on net methylation in situ ......................... 28
3.4.4. Net methylation and demethylation of different Hg pools .................. 30
3.5. Remarks for the future research ................................................................... 31
3.5.1. Determination of Hg chemical speciation on the methylating bacteria
surface proximity .............................................................................................. 31
3.5.2. Experimental approaches for determination of MeHg, and HgII
complexes to LMM thiols in pure bacteria culture ......................................... 33
4.
Conclusions ....................................................................................................... 34
5.
Acknowledgements ........................................................................................... 36
6.
References ......................................................................................................... 38
viii
List of publications
Publications included in this thesis
I.
Determination of Sub-Nanomolar Levels of Low Molecular Mass Thiols in
Natural Waters by Liquid Chromatography Tandem Mass Spectrometry
after Derivatization with p‑(Hydroxymercuri) Benzoate and Online
Preconcentration
Van Liem-Nguyen, Sylvain Bouchet and Erik Björn
Analytical Chemistry, 2015, 87, 1089-1096
II.
Stability constants for mercury (II) complexes with low molecular mass
thiols as determined by competing ligand exchange and liquid
chromatography inductively coupled plasma mass spectrometry
Van Liem-Nguyen, Ulf Skyllberg, Kwangho Nam and Erik Björn
Manuscript in preparation
III. Thermodynamic modelling of the chemical speciation of mercury and
methylmercury under sulfidic conditions in boreal wetland soils
Van Liem-Nguyen, Ulf Skyllberg, and Erik Björn
Manuscript in preparation
IV. Mercury chemical speciation and biological production of low molecular
mass thiols in control of methylmercury formation in boreal wetlands
Van Liem-Nguyen, Ulf Skyllberg and Erik Björn
Manuscript in preparation
V.
Effects of nutrient loading and mercury chemical speciation on the
formation and degradation of methylmercury in estuarine sediment
Van Liem-Nguyen, Sofi Jonsson, Ulf Skyllberg, Mats B Nilsson, Agneta
Andersson, Erik Lundberg and Erik Björn
Under review in Environmental Science & Technology
Paper I is reprinted with permission from Analytical Chemistry. Copyright ©
2015, American Chemistry Society.
ix
Author contribution:
Paper I. The author formulated the scientific approach, led the project planning
work, performed all of the experimental work and made major contributions to
the data evaluation and writing of the paper.
Paper II. The author was largely involved in formulating the scientific research
objectives and approaches, led the project planning work, conducted most of the
experimental work and the data evaluation and was lead author in writing the
manuscript.
Paper III. The author was largely involved in formulating the scientific research
objectives and approaches and in the project planning work, performed most of
the experiments and data processing and made major contributions to writing
the manuscript.
Paper IV. The author was largely involved in formulating the scientific research
objectives and approaches and in the project planning work, performed most of
the experiments and data processing and made major contributions to writing
the manuscript.
Paper V. The author was largely involved in data processing and interpretation
and made major contributions to the writing of the manuscript.
x
Abbreviation
2-MPA
2-Mercaptoproprionic acid
AMA
Advanced mercury analyzer
Aq
Aqueous
Bio
Biological
Chl α
Chlorophyllα
Cys
L-Cysteine
DOC
Dissolved organic carbon
DOM
Dissolved organic matter
EXAFS
Extended X-Ray absorption fine structure
GC
Gas chromatography
Glyc
Monothioglycerol
GSH
Glutathione
GTN
Gästern
HDPE
High density polyethylene
Hg
All chemical forms of mercury
Hg0
Elemental mercury
HgII
Inorganic divalent mercury
IC
Ion chromatography
ICP
Inductively coupled plasma
IQ
Intelligence quotient
K
Stability constant
Ka
Acid dissociation constant
Kd
Distribution between solid/adsorbed and dissolved phases
kd
Methylmercury demethylation rate constant
km
Mercury methylation rate constant
KSN
Kroksjön
LC
Liquid chromatography
LDN
Långedalen
LMM
Low molecular mass
LODs
Limits of detection
xi
MAC
Mercaptoacetic acid
MeHg
All chemical forms of methylmercury
MS
Mass spectrometry
MS/MS
Tandem mass spectrometry
NACCys
N-acetyl cysteine
NOM
Natural organic matter
Org-SRED
Organic reduced sulfur
P.P
Photosynthetic primary biomass production rate
Pen
Penicillamine
PHMB
p‑(hydroxymercuri) benzoate
PP
Polypropylene
R2
Correlation coefficient
RSH
Thiol
RSR
Organic monosulfide
RSSR
Organic disulfides
S XANES
Sulfur X-ray absorption near edge structure
S(-II)
Inorganic sulfide
S0
Elemental sulfur
SD
Standard deviation
sed
Sediment
SKM
Storkälsmyran
SPE
Solid phase extraction
SRB
Sulfate reducing bacteria
STEB
Sodium tetra ethyl borate
SUC
Mercaptosuccinic acid
t48
Time after 48 h from the beginning of the incubation
to
Time at the beginning of the incubation
US EPA
United States Environmental Protection Agency
UV
Ultra violet
wt
Water
β2
Combined stability constant, β2= K1×K2
xii
1. Introduction
1.1. Mercury
Mercury (Hg) is a highly neurotoxic compound, causing negative effects to
humans health and to wildlife. Mercury is a naturally occurring metal in the
Earth's crust with an average concentration of approximately 0.05 mg kg-1. While
the ore of Hg has been found to contain up to 12-14% of naturally occurring Hg, it
exists mainly in the form of cinnabar mineral. 1,
2
Mercury is present in the
atmosphere at low levels (1-170 ng m-3), mostly as elemental mercury (Hgo).3, 4 It
can be released into the atmosphere by natural, e.g. volcanic eruptions, and
anthropogenic activities such as incineration of fossil fuels, mining and industrial
activities.5-7 Globally, industrialization have significantly increased anthropogenic
Hg release by a factor of 3 to 5 and have increased chronic Hg exposure to
humans as well as aquatic and terrestrial organisms.5,
8
Despite societal
undertakings, Hg pollution is still a major environmental issue globally. 9, 10
Exposure to high Hg concentration can damage the central nervous system and
in worst case cause death.11, 12 This mostly happens for those who work in mines
under poor conditions and limited knowledge of health and safety measures.3
Consumption of fish and rice is the main source of Hg in humans.13-16 The first
officially discovered Hg pollution incident was in Minamata city, Japan’s Kyushu
Island in May 1956.11 This led to the discovery of the “Minamata disease”; a
disease of methylmercury (MeHg) poisoning through the consumption of fish
and shellfish.11 In March 1992, 2252 patients were officially diagnosed of
Minamata disease and 1043 fatalities was recorded.12
In Sweden, Hg levels in piscivorous fish greatly exceed the levels (0.02 mg Hg kg1
fresh weight (EU 2008))17 for environmental monitoring target in most of the
hundred thousand lakes.18 The Hg concentration in fish 0.2-0.5 mg Hg kg-1 (wet
weight) is considered as safety level for fish consumption according to United
Nations Environmental Programme (UNEP).19 About 7 % of women of childbearing age in the USA are estimated to have consumed food with Hg
concentrations that exceeds health guideline.3. A recent study concluded that
about 1.8 million children are born every year in Europe with Hg concentration
exceeding the limit of 0.58 µg g-1 (in hair), which has negative effects on the IQ
1
and that the cost for the society of Hg exposure is 9000 million € per year.20 In
Sweden, around 10% of children are born with Hg concentration above the limit
(0.58 µg g-1, in hair), and the cost for Hg exposure is 150 million € per year for
the Swedish society.20
Natural events and anthropogenic activities emit mostly inorganic forms of Hg,
but the organic form, methylmercury (MeHg), is the dominant in fish and rice.16,
21, 22
Methylmercury is much more toxic than inorganic mercury thus in
combination with its persistence and bioaccumulating properties the formation
of MeHg in the environment is of major importance.
1.2. Chemical species and chemical speciation of mercury
According to IUPAC, chemical species is defined as “specific form of an element
defined as to isotopic composition, electronic or oxidation state, and/or complex
or molecular structure” and the chemical speciation is defined as “distribution of
an element amongst defined chemical species in a system”.23 Therefore, the
chemical speciation of Hg with respect to redox is the distribution amongst
divalent (HgII), monovalent (HgI) and elemental mercury (Hgo). The chemical
speciation with respect to molecular complexes composition of inorganic divalent
Hg (HgII) is the distribution of different HgII complexes, e.g. Hg(SH)2, HgCl2 and
Hg(SR)2, in for instance wetland system. Also the term “geochemical pools” of
HgII and MeHg is relevant in this thesis and was defined by Jonsson et al.:
“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”.24
In the solid/adsorbed phase HgII is mostly present as cinnabar (HgS(s)) form or
bound to thiol (RSH) functional groups of NOM. Inorganic divalent mercury is
also present in the aqueous phase as strong complexes with soft Lewis bases
(ligands) such as inorganic sulfide, thiols and halides (but not fluoride). 25,
26Typically,
HgII complexes are low coordinated. Two-coordinated complexes are
most common followed by three- and four-coordinated.27-30 The linear twocoordination of HgII is caused by relativistic effects and is a very unusual
structure for a metal.31 On the other hand, the strong covalent character of the C–
Hg bond is the reason for the high stability of the MeHg molecule and MeHg also
2
has similar characteristics as HgII when it comes to the strong complexation with
soft Lewis bases.28, 32, 33
Reports on complete chemical speciation of HgII in natural environments,
covering both solid/adsorbed and aqueous phases are quite few due to the lack of
analytical methods to directly measure specific relevant Hg II species at natural
concentrations, and the need to rely on thermodynamic modeling. Most studies
focus on the chemical speciation of HgII in either soil34 or aqueous (dissolved)
phases.35,
36
One reason for this is a significant uncertainty in the stability
constants for HgII complexes formed by ligands such as low molecular mass
(LMM) thiols, thiols associated with natural organic matter (NOM-RSH) and
inorganic sulfides (S(-II)).25,
37
In this work, comprehensive thermodynamic
models were constructed for MeHg and HgII, including both the solid/adsorbed
and dissolved phases. The models were evaluated by comparing the model results
with actual measurements of concentrations of MeHg, Hg II and important
ligands (particularly LMM thiols) in soil and porewater of the boreal wetlands.
Such an evaluation has not been done in these environments before. Even though
LMM thiols are present at low concentrations in environments they are believed
to play an important role in transformations of HgII.38-40
1.3. Methylation of HgII
The transformation of HgII to MeHg (methylation) is mainly the result of biotic
processes under suboxic and anoxic conditions. Sulfate reducing bacteria (SRB)
have been identified as important methylators.41,
bacteria
(IRB),43, 44
methanogens,45
42
Recently, iron reducing
and other types of microbes46 have been
reported producing MeHg in natural environments. The methylation rates have
been suggested to be limited by the amount of HgII available for uptake by
methylating
bacteria,
which
is
controlled
by
inorganic
sulfide
(S(-II))
concentration, natural organic matter (NOM) functional groups, and pH. 47 Due
to the strong affinity of HgII to reduced sulfur, the chemical speciation of HgII is
mostly driven by the relative proportions of S(-II) and thiol (RSH) groups.25,
47Inorganic
divalent mercury forms complexes with RSH groups in a two-
coordinated structure, e.g.Hg(SR)2.27 In addition, different geochemical pools of
HgII have also been demonstrated to control MeHg production. 48 The presence of
3
mackinawite FeS(s) has been shown to decrease MeHg production in wetland
slurries due to reduced availability of HgII for methylation.49, 50
It has been shown that Hg(SH)20(aq) and HgII complexed by specific low
molecular mass (LMM) thiols i.e. Hg(LMM-RS)2 are available for uptake by
methylating
bacteria.36,
38
A
recently
developed
theory
suggests
that
nanoparticulate HgS(s) is indirectly or directly taken up by methylating bacteria,
yet the mechanism for the uptake remains to be identified.51 In natural waters,
sediments, and wetland environments concentrations of LMM thiols and S(-II)
vary largely. Low molecular mass thiols normally range from sub nM levels
(LODs) to hundreds nM [Paper I and Manuscript III & IV], but in some cases
they may reach µM52, 53 or even mM54 concentrations. Under suboxic to anoxic
conditions S(-II) concentrations may vary from 20 nM (LOD)55 to mM.56,
57
Therefore, concentrations of bioavailable HgII complexes may vary substantially
depending on the concentrations and molar ratio of LMM thiols and S(-II). The
rate of HgII methylation has been shown about 5–6 times higher in the presence
of specific LMM thiols e.g. Cys or MAC, as compared to the presence of similar
concentration of S(-II).38 Thus, it is of utmost importance to develop relevant
thermodynamic models to predict at which molar ratio between S(-II) and LMM
thiols, the uptake by methylaing bacteria is mainly driven by Hg II–sulfides or by
HgII–LMM thiols complexes.
Wetlands have seasonally variable water regimes covering oxic-suboxic-anoxic
conditions. These frequent seasonal fluctuations turn wetlands into dynamic
redox environments and often enhance MeHg formation. This makes wetlands
become critical sources of MeHg release, and very important environments for
Hg cycling.58, 59
Geochemical pools of Hg have been assumed to be different with respect to their
availability for methylation, bioaccumulation, transportation and mobilization.47,
60.
Jonsson et al. reported a higher net MeHg formation of Hg II newly deposited
to sediment, as compared to the aged HgII sediment pool, by use of a mesocosm
approach.48 In this thesis work, a similar observation of differentiated availability
of HgII geochemical pools for methylation was extended by investigating the effect
of nutrient loading on rates of methylation and demethylation.
4
1.4. Aims of the thesis
The aims of this thesis are:
(1) To determine LMM thiols concentration in natural waters as well as to
establish stability constants for different HgII–LMM thiols complexes by
developing and apply novel analytical methods (Paper I & II).
(2) To determine the concentration of important MeHg and HgII species in soils
and porewaters by establishing comprehensive thermodynamic speciation
models for MeHg and HgII (Paper III).
(3) To investigate if the chemical speciation of Hg II control MeHg formation in
boreal wetlands and estuarine sediment environments and control the effect of
nutrient loading on HgII methylation (Paper IV &V).
5
2. Materials and experimental approaches
2.1. Study sites
2.1.1. Boreal wetlands
Boreal wetlands have been reported as sources of MeHg formation and
transportation.56,
58
Detailed information on the characteristics of the wetlands
selected for the study are found in Paper III and in Tjerngren et al.58 Briefly, four
different boreal wetlands were selected and studied: two in southern and two in
northern Sweden. The southern sites were Långedalen (LDN) and Gästern
(GTN), and Storkälsmyran (SKM) and Kroksjön (KSN) represent the northern
sites. The map and location of sampling spots are shown in Figure 1 and Table 1.
Figure 1. Maps showing sites and sampling locations (black filled squares). Black stars
indicate locations of small dams, and black solid lines denote streams.
6
Table 1. Sampling locations and properties of the boreal wetlands (information from
Tjerngren et al., 2012).58
Location
Wetlands
Long (N)
Lat (E)
Central wetland
type
SKM
63°57'52.0"
20°38'47.0" Riparian
KSN
63°56'41.6"
20°38'8.7"
LDN
58°19'45.7"
12°29'8.1" Fen/bog
GTN
57°27'18.8" 16°35'0.4"
Dystrophic lakepeatland complex
Mesotrophic lakepeatland complex
Catchment
area (km2)
Central
wetland area
(ha) and (%)
Donating vegetation
Group
0.48
0.020 (4.2)
Carex spp., Sphagnum spp.,
Polytrichum spp.
1.0
0.26 (26)
Carex spp., Sphagnum spp.
2.0
0.106 (11)
Scirpus spp., Carex spp.,
Peatland nutrient
Sphagnum spp., broad-leaved
gradient
grasses, Calluna vulgaris.
23
0.58 (2.5)
Phragmites australis, thypa
latifolia, Sphagnum spp.
Northern, nutrient
poor peatlands
Northern, nutrient
poor peatlands
Southern, nutrient
rich wetlands
2.1.2. Coastal marine system
A marine sediment was sampled from Öre river estuary. The estuary is located in
the Bothnian Sea (site 63° 33.905′ N, 19° 50.898′ E) at the Swedish east coast
and covers an area of 50 km2 with a total water volume of 109 m3 and an average
depth of 16.4 m. The sediment has a total Hg concentration of approximately 200
pmol g-1 (d.w) which means it is classified as non-contaminated (background site)
sediment.
2.2. Experimental approaches
2.2.1. Determination of LMM thiols in natural waters
Thiols can be associated with natural organic matter (NOM) or low molecular
mass organic molecules and are present in so-called thiol (RSH) or organic
disulfides (RSSR) forms. Previous studies mostly reported LMM thiols
concentration as the sum of RSH and RSSR forms after adding a reducing agent
to convert RSSR to RSH prior to measurement.61, 62 However, the affinity of the
RSH is much stronger for most metals (including HgII) as compared to the RSSR
form. It is therefore important to determine the concentration of RSH separately
from RSSR. In Paper I, we developed a novel analytical method to determine 16
different LMM thiols using solid phase extraction online pre-concentration
hyphenated to liquid chromatography tandem mass spectrometry (SPE/LCMS/MS). The method allows determination of LMM thiols in natural waters at
sub nM levels of limits of detection (LODs).
7
To increase the sensitivity, selectivity and preservation of LMM thiols,
p‑(Hydroxymercuri) Benzoate (PHMB) was used as a probe. The PHMB probe is
selective and has strong affinity for the RSH groups. It is however an extremely
toxic compound,61 and therefore its concentration should be optimized for the
derivatization step. A simplified version of the optimization (Paper I) suggests
that the minimal concentration of PHMB for the derivatization could be
calculated by the following equation:
[PHMB] = 3×([NOM-RSH(aq)] + [S(-II)])
The pH of the boreal wetland porewaters is relatively low, ranging from 4 to 6.
Consequently, inorganic sulfide (S(-II)) is mostly present as H2S(aq, g), which was
evaporated during sample preparation. When dealing with samples containing
high (S-II) concentration at neutral or alkaline pH (pH>7), for instance sediment
porewater, an acidification step can be used to remove S(-II).
2.2.2. Determination of total thiols in natural waters and soils
Total thiol (LMM thiols and thiol associated with NOM) in waters and soils was
determined using Sulfur K-edge X-ray absorption near edge structure
spectroscopy (S K-edge XANES). For porewater samples, one liter was filtered
through a 0.22 µm membrane filter in a glove box filled with N 2 gas. This sample
was then frozen at -20°C and freeze-dried. As illustrated by the spectrum in
Figure S3, Paper III, one peak of reduced organic sulfur and one peak of oxidized
sulfur was obtained. The latter peak was highly dominated by sulfate (SO 42-) in
the porewater. The reduced sulfur peak includes the signal of the three organic
sulfur forms: thiol (RSH), organic disulfides (RSSR) and organic monosulfide
(RSR). The absorption energies of those three groups of compounds are close to
each other and partly overlapping, and therefore often summed up as organic
reduced sulfur (Org-SRED).63 The concentration of SO42- in porewater was also
determined by ion chromatography (IC). Based on the measured concentration of
SO42- by IC and the relative contribution of Org-SRED and SO42- to the S XANES
spectrum, the absolute concentration of Org-SRED was determined.
Prior to S XANES analyses of soil samples, wet soils were freeze-dried and
homogenized. The total sulfur in soil samples was determined by LECO analyzer
8
whereas elemental sulfur (S0), extracted from fresh soils, were determined by
liquid chromatography with UV adsorption.64 The obtained S XANES spectra
were deconvoluted into absolute concentrations of Org-SRED, sulfoxide, sulfonate
and sulfate based on the relative normalized XANES signal of the respective
peaks for these S functional groups and the total sulfur concentration in the
sample (Figure S3, Paper III). Finally, the concentration of RSH groups was
calculated as 30% of Org-SRED, based on previously combined Hg LIII-edge
EXAFS and S K-edge XANES determinations of HgII bonding to RSH groups in
NOM from similar types of boreal soils and waters.27, 63
2.2.3. Determination of total Hg and MeHg in natural waters and soils
Total Hg in waters was analyzed by isotope dilution analysis (IDA) combined
with mercury cold vapor generation ICPMS while the analysis of total Hg in soils
was performed by AMA 254 (LECO). Total MeHg in waters was analyzed by IDA
combined with thermal desorption gas chromatography (GC) ICPMS after
derivatizing with sodium tetraethyl borate (STEB).4 Prior to analysis, soil samples
were spiked enriched MeHg isotope for isotope dilution analysis (IDA). Following
the samples were homogenized and MeHg in soils were extracted using
KBr/CuSO4/H2SO4/CH2Cl2 and determined by GC-ICPMS after derivatizing with
STEB.65
2.2.4. Determination of the methylation and demethylation rate
constants
The methylation rate constant of HgII (km) and MeHg demethylation rate
constant (kd) in soils and sediments were determined using enriched stable
isotope tracers. Both km and kd were described by irreversible pseudo first-order
kinetic models.66 Experimentally, approximately 10.00 g (with 2 decimal digits)
of wet soil/sediment samples were weighed into two sample sets inside the glove
box. Dissolved Me199HgCl(aq) and
204Hg(NO
3)2(aq)
(and/or
200HgII-NOM(ads)
adsorbed tracer), corresponding to 10-30 % of the total MeHg and HgII
concentrations in the samples were added. The synthesis of
200HgII-NOM(ads)
tracer was described by Jonsson et al.67 based on Skyllberg and Drott.68 One
sample set was immediately frozen at -20oC, designated to, while the other sample
set was incubated in a N 2 filled glovebox at room temperature in the dark for 48
9
h, designated t48. The incubation was terminated by placing the t 48 sample set
into a freezer. The determination of MeHg in the incubation study was carried
out with the same procedure for the determination of MeHg in soil/sediment.
The km and kd were calculated using the equations below.
[Me204Hg]t48 – [Me204Hg]t0
km =
kd =
204
II
(d-1)
[ Hg ]addition×2
-1×(ln[MeHg199Hg]t48 – ln[Me199Hg]t0)
2
(d-1)
2.2.5. Determination of stability constant for Hg(LMM-RS)2
complexes
The stability constants f0r 15 Hg(LMM-RS)2 complexes were determined using a
competing ligand exchange method combined with liquid chromatography
ICPMS. In the first step, iodide (I-) ligand was used as the competing ligand to
determine the stability constant for Hg(MAC)2 and Hg(2-MPA)2 due to wellknown stability constants for the complexation of Hg II with I-.69 In the second
step, to enable determination of LMM thiol ligands having different retention
time in the LC column, either MAC or 2-MPA was used as a competing ligand
when establishing the stability constant for the 13 other Hg(LMM-RS)2
complexes.
2.2.6. Chemical speciation of MeHg and HgII in boreal wetlands
Concentrations of MeHg, HgII and important ligands for the determination of
MeHg and HgII chemical speciation in soils and porewaters were measured.
Those parameters were then used as input to the WinSGW software to calculate
the concentration of MeHg and HgII complexes pertaining to solid/adsorbed and
dissolved phases. The reactions (1) to (7) were the most significant reactions with
strong influence on the chemical speciation of MeHg and Hg II in the wetland
system.
MeHg+ + H2S = MeHgSH
+ H+
MeHg+ + LMM-RSH = MeHgSR-LMM + H+
(1)
(2)
10
MeHg+ + NOM-RSH = MeHgSR-NOM + H+
(3)
Hg2+ + 2H2S = Hg(SH)2+ 2H+
(4)
Hg2+ + H2S = HgS(s) + 2H+
(5)
Hg2+ + 2LMM-RSH = Hg(LMM-RS)2+ 2H+
(6)
Hg2+ + 2NOM-RSH = Hg(NOM-RS)2+ 2H+
(7)
2.2.7. Mesocosm experiment
Mesocosm scale experiments are a powerful tool for simulating natural
environmental conditions and to control factors which affect, in our case, MeHg
formation in marine ecosystems. Our mesocosm experiments were conducted in
6 double-mantled high density polyethylene (HDPE) tanks (5 m × 0.75 m ⌀),
located at the Umeå Marine Sciences Centre. Intact sediment cores were
manually sampled from the Öre river estuary by divers at the site coordinates 63°
33.905′ N, 19° 50.898′ E and at water depth of 5-7 m using custom-made
sampling devices, then inserted into the mesocosm tanks. Solid/adsorbed
mercury tracers were injected into the sediments and dissolved mercury tracers
were added to the water columns. The tanks were temperature controlled at
different heights via an outer glycol layer. The temperature of the water column
was controlled in two sections with 15oC in the upper part (above 3.5 m) and 10oC
in the lower part (below 3.5 m). The convection of the upper part of the water
column was obtained by purging with ~20 mL s -1 air at 3.2 m from the water
surface. 150 W metal halogen lamps (Master color CDM-T 150w/942 G12 1CT)
was used as light sources with 12/12 hours on/off cycles.
The three Hg sediment tracers i.e.
200HgII-NOM
sed,
β-201HgSsed and Me198Hg-
NOMsed were injected to the sediments at 1 cm depth prior to loading of the
brackish water to the tanks. The other two Hg water tracers,
Me199Hgwt,
204HgII
wt
and
were then added to the water columns. Nutrients (N, P) were also
added to the water columns from moderate to rich nutrient level. Total
experimental time was 30 days.
2.2.8. Mercury isotope tracers and nutrient preparation
The Hg enriched
201Hg
196Hg
(98.11%) and
(50%),
204Hg
198Hg
(92.78%),
199Hg
(91.95%),
200Hg
(96.41%),
(98.11%) (as HgO or HgCl2) were purchased from Oak
11
Ridge National Laboratory (TN, USA). The
prepared by diluting
201HgS
sed
204HgII(aq)
204HgII
wt
and Me199Hgwt tracers were
and MeHg199Hg(aq) in MQ water. The β-
and isotopically enriched MeHg tracers were synthesized as described
by Jonsson et al.67 and Snell et al.70 The
tracers were prepared by adding
200HgII-NOM
200HgII(aq)
sed
and Me198Hg-NOMsed
and Me198Hg(aq) to a homogenized
organic soil previously characterized by Skyllberg and Drott. 68 The Hg/RSH
molar ratio was kept in the range of 1:2 HgII:RSH and 1:1 MeHg:RSH complex
stoichiometry.68 Nitrate (NO3-), phosphate (PO43-) and ammonium (NH4+)
solutions were prepared from salts of NaNO3, NaH2PO4×H2O and NH4Cl,
respectively.
12
3. Results and Discussion
The presence of LMM thiols has, in pure bacteria culture, been demonstrated to
enhance methylation rates of HgII.
38, 40
However, the characterization of LMM
thiols in natural environments has so far been very limited due to a lack of
powerful (high accuracy, low detection limits) analytical methods. Therefore, in
Paper I, we focused on the development of a new analytical method for
determination of LMM thiols in natural environments. Complexes of HgII with
LMM thiols are believed to be present mainly as two-coordinated monodentate
complexes i.e. Hg(LMM-RS)2, and to be taken up by methylating bacteria. 27, 29, 30,
38
Many different types of LMM thiols occur in natural environments39, 52, 53 and
their effects on HgII methylation rates are different.38 In the literature, not all
reported LMM thiols have established stability constants for complex with HgII,
and some reported constants also vary substantially. Hence, the aim of Paper II is
to fill a gap of knowledge and to establish stability constants for an
environmentally important group of Hg(LMM-RS)2 complexes (15 complexes).
Paper I and II are essentially analytical chemistry oriented laboratory studies. In
a contrast, Papers III, IV and V were more focusing on applying the methods and
knowledge generated in the laboratory experiments to study processes in natural
environments. The chemical speciation of MeHg and HgII in boreal wetlands was
determined by the establishment of comprehensive thermodynamic models for
MeHg and HgII in Paper III. The relation between HgII species in porewaters and
methylation as well as factors in control of the methylation were addressed in
Paper IV. By the use of a mesocosm setup, Paper V investigated the methylation
of different HgII sediment pools, and dissolved and solid species of HgII. Further
more, the effects of nutrient loading on MeHg formation was also investigated. It
demonstrated the roles of HgII availability and methylating bacteria activity for
the methylation. The scheme in Figure 2 shows the contribution of each Paper in
addressing MeHg formation
13
Paper I
Paper II
Log k Hg(LMM-RS)2
LMM thiols
MeHg
formation
HgII species in
boreal wetlands
*Solid/absorbed phase of HgII
*Effect of nutrient loading
Paper III
Paper V
Relation of HgII species in
porewaters and MeHg
formation
Paper IV
Figure 2. Schematic aims for each Paper of the thesis work
3.1. Determination of LMM thiols in boreal wetland
porewaters (Paper I)
Low molecular mass (LMM) thiols are a diverse group of compounds, which may
play important roles in aquatic ecosystems even at nM concentrations.52,
53, 71
Methods for determination of LMM thiols are mostly dedicated for biological
samples.61,
62, 72
Methods based on liquid chromatography (LC) separation with
UV adsorption73,
74
and fluorescence detection75-78 are usually constrained by
relatively high limits of detection (LODs) and auto-quenching, respectively. With
mass spectrometry, the selectivity and sensitivity for LMM thiols determination
are improved. Rao et al.61 and Bakirdere et al.62 reported LODs from 1 to 32 nM
for 6 LMM thiols determined by high resolution Orbitrap mass spectrometry.
Still, the LODs of those methods are not low enough to cope with natural water
matrices. Because of these shortcomings, we developed an analytical method for
the determination of LMM thiols in natural waters (can also be applied for
biological samples) using SPE online pre-concentration liquid chromatography
tandem mass spectrometry (SPE/LC-MS/MS).
14
3.1.1. Limits of detection of the method
By the use of SPE online pre-concentration, the analysis time and running costs
were highly reduced when compared to conventional SPE offline preconcentration method. The LODs of 16 investigated LMM thiols ranged from
0.06 nM to 0.5 nM (Table 3, Paper I). The LOD of each LMM thiol substantially
depends on both its molecular mass and hydrophobicity. LMM thiols with higher
molecular mass and stronger retention on reversed phased LC column (i.e. more
hydrophobic) showed better LODs. This could be explained by the recovery of
PHMB-thiol complexes on the HLB oasis SPE online cartridge and the speciesspecific sensitivity of electrospray ionization MS. The SPE online column covers
mixed mode (lipophilic and hydrophilic) interactions but compounds with high
hydrophobicity show strong interaction and high recovery on the SPE column.
For electrospray ionization MS, a high signal-to-noise-ratio was most likely
achieved for PHMB-thiol complexes with high molecular mass. By combining the
two characteristics, N-acetyl-penicillamine (the second largest molecular mass,
191 Da, and the most hydrophobic) and mercaptoethanol (the smallest molecular
mass, 78 Da) showed the best (0.06 nM) and worst (0.5 nM) LOD, respectively
(Table 3, Paper I).
3.1.2. Determination of LMM thiols in boreal wetland porewaters
In total eight different LMM thiols were detected in the boreal wetland
porewaters, with concentrations in nM range (Table 2). For individual LMM
thiol, concentration of separate RSH form ranged from 0.1-77 nM whereas the
sum of RSH and RSSR forms gave a range of 0.7-110 nM. Low molecular mass
thiols such as Cys, NACCys and GSH are known to be of direct biological origin
(biomolecules) while MAC, 2-MPA, Glyc, Pen and SUC have an indirect biological
origin, can be formed by addition of sulfide to unsaturated alkyl chain of organic
compounds.73, 79-81 Average concentrations of total LMM thiols and reduced LMM
thiols in the porewaters of all four studied wetlands were 41 nM and 22 nM,
respectively. Concentrations of LMM thiols were highest at the northern lakewetland site (KSN), and the bog-fen peatland gradient (LDN), with average total
LMM thiols concentration of 50 nM. At the northern riparian wetland site (SKM),
the average concentration was 40 nM, and 20 nM for the mesotrophic southern
15
wetland (GTN). At the KSN site, Glyc was the most abundant, contributing more
than 90% of total LMM thiols. At LDN, SKM and GTN sites, Cys and MAC were
the most abundant thiols. A positive relationship could be established between
the concentration of LMM thiols of direct biological origin and rates of HgII
methylation.
Table 2. The concentration of Σ(disulfides (RSSR) and thiol (RSH) forms), and separate
RSH form in parenthesis for LMM thiols in the boreal wetland porewaters.
Samples
Σ RSSR and RSH, and (RSH) concentration (nM)
MAC
SKM-1
SKM-2
SKM-3
SKM-4
SKM-5
SKM-6
SKM-7
SKM-8
KSN-1
KSN-2
KSN-3
KSN-4
KSN-5
KSN-6
KSN-7
KSN-8
LDN-1
LDN-2
LDN-3
LDN-4
LDN-5
LDN-6
LDN-7
LDN-8
LDN-9
LDN-10
GTN-1
GTN-2
GTN-3
GTN-4
GTN-5
Glyc
19 (14)
29 (23)
40 (15)
35 (8.9)
38 (20)
22 (14)
29 (18)
5.3 (2.3)
7.1 (3.1)
6.0 (2.4)
9.0 (3.2)
28 (16)
62 (32)
51 (38)
38 (30)
12 (9.2)
16 (13)
14 (8.5)
10 (6.9)
6.8 (5.0)
6.5 (2.1)
10 (3.0)
9.2 (5.4)
15 (3.2)
15 (10)
20 (12)
50 (30)
74 (45)
47 (31)
28 (10)
22 (16)
21 (12)
26 (20)
110 (77)
Cys
7.6 (2.7)
4.8 (2.0)
4.5 (2.2)
5.1 (1.8)
3.1 (1.2)
12 (4.3)
3.5 (3.2)
9.2 (4.6)
2.9 (1.1)
3.2 (1.0)
2.5
1.5
2.8 (1.2)
3.4
2.8
9.1 (3.1)
8.0 (2.4)
6.3 (2.6)
6.6 (2.6)
2.1 (1.5)
2.9
2.5 (2)
3.2
2.8
1.5
2.3 (0.92)
3.1 (1.2)
5.2 (1.6)
3.3 (0.92)
2.5 (1.1)
Pen
SUC
NACCys
GSH
2-MPA
0.59
5.7 (1.5)
2.0
0.55
7.8 (3.9)
1.0
2.5 (1.2)
3.2
4.6
0.86 (0.24)
0.59 (0.22)
0.8 (0.3)
1.0 (0.4)
0.62 (0.25)
0.78
11 (2.8)
8.7 (2.3)
0.73 (0.38)
0.86 (0.50)
1.2 (0.87)
0.70 (0.17)
1.6 (0.21)
1.2 (0.47)
1.0 (0.63)
11 (2.2)
15 (2.8)
2.8
0.88
0.51
2 (1.5)
15 (2.1)
1.5 (1.5)
0.73
0.42
0.52 (0.35)
0.71 (0.45)
0.71 (0.58)
1.8 (0.80)
0.64 (0.43)
0.84 (0.25)
0.6
2.8 (1.2)
3.2
Total
27 (17)
34 (25)
54 (19)
47 (11)
3.1 (1.2)
62 (30)
26 (17)
39 (23)
56 (33)
85 (50)
49 (31)
29 (10)
22 (16)
30 (16)
30 (21)
120 (84)
61 (24)
100 (40)
70 (46)
47 (34)
15 (11)
21 (15)
18 (12)
13 (6.9)
11 (5.0)
8.5 (2.5)
13 (4.3)
13 (7.2)
25 (5.6)
19 (11)
23 (13)
3.2. Determination of stability constants for Hg(LMMRS)2 complexes (Paper II)
It is a challenge to determine stability constants for the very stable HgII–LMM
thiol complexes. The wide range of stability constants reported for these
complexes before 1980 was extensively discussed at by Casas and Jones.82 By
applying an direct determination of very low concentration of free Hg2+ ions, they
reported the log β2 for Hg(Cys)2, Hg(Pen)2 and Hg(MAC)2 to be in the range of
16
40.0 to 45.0.82, 83 In later works, addition of a competing ligands such as iodide
(I-) or bromide (Br-) ion or the lipophilic thiol dithizone shifted the HgII
speciation towards a separable, measurable and identifiable entity.30,
84-86
Even
with this approach, there are still uncertainties in stability constants reported, as
exemplified by log β2 for Hg(Cys)2 complex, ranging between 38.230 and 43.586.
Therefore, there is still a substantial uncertainty remaining before a consensus on
stability constants for HgII complexes with LMM thiols can be reached. It has
been established that Hg(LMM-RS)2, Hg(LMM-RS)3, and Hg(LMM-RS)4
complexes may all form depending on pH and the HgII/RSH molar ratio.29, 30 At a
pH lower than 6, Hg(LMM-RS)2 is the most predominant form, while the
proportions of Hg(LMM-RS)3 and Hg(LMM-RS)4 increase at higher pH as
demonstrated for the complexation of HgII with Cys, Pen and NACCys.29, 30, 87, 88
For larger LMM thiols like GSH, the formation of higher coordination numbers
than two is less common. At pH of 7.4, Hg(GSH)2 contributes 95% of all HgII
complexes.29 This could be explained by steric hindrance effects of large ligand
molecules, limiting the coordination number of the complex. In line with this,
spectroscopic methods such as EXAFS and NMR indicate a high dominance of
the Hg(NOM-RS)2 structure in the complexation of HgII with NOM associated
RSH (NOM-RSH) in soil and natural water at pH below 7.27, 30
The concentration of Hg(LMM-RS)2 complexes in the environment is estimated
to be extremely low, normally below fM range, in porewaters.89 Current analytical
methodologies do not allow direct detection of such complexes. Therefore, the
determination of Hg(LMM-RS)2 still needs to rely on thermodynamic modeling
and correct stability constants for Hg(LMM-RS)2 complexes are obviously
crucial. In this study, we determined stability constants for 15 different Hg(LMMRS)2 complexes using a novel analytical method based on competing ligand
exchange (CLE), liquid chromatography separation and ICPMS detection.
Determined stability constants for the 15 investigated complexes are reported in
Table 3. When writing the reaction without considering the protonation of RSH
group (reaction 1a) the log K of Hg(LMM-RS)2 range from the lowest value of
19.6 to the highest value of 21.0 corresponding to Hg(Cyst) 2 and Hg(2-MPA)2,
respectively. For the reaction 1b (Table 3), where the pKa value of the RSH
groups is included in the reaction (log β2= log K + 2pKa), the log β2 showed a
17
wider range from the lowest value of 34.6 for Hg(CysGly)2 to highest value of 42.1
for Hg(3-MPA)2. The pKa value of the RSH group is important for the
determination of β2, they were thus calculated by the Atomic Charge model
(quantum mechanics model) reported by Ugur et al.90 as shown in Paper II.
Table 3. Stability constants for Hg(LMM-RS)2 complexes reported in agreement with
reaction 1a and 1b and pKa values for RSH groups (reaction 2).Complexes are sorted from
highest to lowest log K for reaction (1a) at an ion strength of zero (I=0) M and pH=3.0.
Complexes
Stability constant (± SD)
log K (1a)
log β 2 (1b)
pKa (2)
Hg(2-MPA)2
21.0
41.5 (±0.1)
10.27
Hg(NACPen)2
20.9
40.1 (±0.2)
9.59
Hg(NACCys)2
20.6
40.2 (±0.2)
9.79
Hg(SUC)2
20.6
41.7 (±0.2)
10.55
Hg(3-MPA)2
20.6
42.1 (±0.2)
10.76
Hg(MAC)2
20.6
40.9 (±0.2)
10.16
Hg(Glyc)2
20.6
39.4 (±0.2)
9.42
Hg(GluCys)2
20.5
40.3 (±0.3)
9.92
Hg(ETH)2
20.5
40.3 (±0.2)
9.89
Hg(Pen)2
20.3
36.9 (±0.2)
8.30
Hg(GSH)2
20.2
38.8 (±0.2)
9.28
Hg(Cys)2
20.1
37.5 (±0.2)
8.64
Hg(CysGly)2
19.9
34.6 (±0.3)
7.34
Hg(HCys)2
19.7
39.4 (±0.2)
9.87
Hg(Cyst)2
19.6
40.3 (±0.2)
10.33
(1a) Hg2+ + 2RSH = Hg(SR)2 + 2H+
(1b) Hg2+ + 2RS- = Hg(SR)2
(2) RSH = RS- + H+
3.3. Mercury chemical speciation in the boreal wetlands
(Paper III)
The biogeochemistry of Hg in the environment is complex. A multitude of
parameters can affect the distribution of Hg species, control their solubility and
subsequent availability for methylation, demethylation and bioaccumulation.48, 63,
68, 89, 91
18
In the 1990s Morel et al.92 gave an overview of the chemical speciation of Hg in
aquatic and terrestrial environments but they did not include the role of LMM
thiols and NOM-RSH. Later studies revealed the important role of NOM and
elucidated that Hg has a strong bond with NOM via complexation with RSH
groups.27, 63, 93-95 Hitherto, there are very few studies including LMM thiols in Hg
chemical speciation due to lack of stability constants for HgII–LMM thiol
complexes and relevant analytical methods for determination of LMM thiols
concentration in natural environments. Although LMM thiols are present at a low
concentration of sub nM to 200 nM39 (in some cases can up to µM52, 53 and mM54
ranges) in natural environments, they are however believed to play an important
role in transformation of Hg.38-40, 54
3.3.1. Geochemical properties at the wetland sites
The boreal wetland porewaters examined in this study showed a typically low pH,
ranging from 4.17 to 6.06 with an average of 5.01. Due to the low pH of the
porewaters, H2S(aq, g) was the dominant species of inorganic sulfide (S (-II)), with
a concentration range from below limit of detection (LOD = 0.3 µM) to 51 µM.
The average concentration of HgII in soils was similar for the three wetlands
SKM, KSN and GTN, about 470 nmol kg-1, and almost double at the fourth site:
LDN, with an average of 900 nmol kg-1. The average MeHg concentrations were
highest at LDN 90 nmol kg-1 , followed by SKM 47 nmol kg-1, KSN 19 nmol kg-1,
and GTN 10 nmol kg-1. The NOM-RSH(aq) concentration in the porewaters
ranged from 0.84 µM to 9.1 µM with an average of 3.8 µM. The sum of
concentrations of the 8 detected LMM thiols (MAC, Cys, Pen, NACCys, GSH,
Glyc, 2-MPA and SUC) ranged between 1.2 nM to 84 nM with an average of 22
nM. In the solid/adsorbed phase of soils, the concentration of NOM-RSH(ads)
was about 800 times higher than NOM-RSH(aq) in the aqueous phase.
Elemental sulfur (S0) was detected in all soil samples, with an increasing average
concentration in the order KSN (12 µmol kg-1) <GTN (45 µmol kg-1) <SKM (70
µmol kg-1) <LDN (150 µmol kg-1).
3.3.2. Chemical speciation model for MeHg
The input for MeHg chemical speciation model included pH, concentrations of
Cl-, HgII, MeHg, NOM-RSH and LMM thiols, S0 and H2S(aq) in soil and
19
porewater compartments. The fit of MeHg model was tested by comparing
measured and modeled concentrations of MeHg in porewaters. In the literature,
stability constants for the formation of MeHgSR-NOM complexes were reported
to vary between 15.6 and 17.5 and pKa of the RSH group to vary between 8.5 and
10.18.32, 33, 63 Therefore, during fitting of our model, the log K of MeHgSR-NOM
complex and the pKa of RSH group were allowed to vary between 16.0 and 17.5
and between 8.5 and 10.0, respectively. The best fit between measured and
modeled concentration of MeHg in the dissolved phase was achieved with a log K
of MeHgSR-NOM of 17.5 and a pKa of 9.0 for the RSH group as shown in Figure
3. The linear relationship between modeled and measured data had a slope of
0.77 (ideally it should be 1.0) and a coefficient-of-determination (R2) of 0.75
(n=31), which can be considered an excellent fit considering uncertainties in
measurements and lack of true equilibrium in natural systems.
Due to high abundance and strong affinity for RSH groups, more than 99% of
MeHg was predicted to bind to thiol associated with NOM i.e. MeHgSRNOM(ads) in the solid/adsorbed phase, with a log concentration of -8.5±0.5 M.
In porewater, the average concentration of MeHgSR-NOM(aq) was ten times
higher than the average concentration of MeHgSH0(aq), with log concentration of
-11.4±0.4 M and -12.3±0.9 M, respectively. Due to low concentration of LMM
thiols as compared to NOM-RSH(aq), the concentration of MeHgSR-LMM(aq)
was about a factor of 100 lower than MeHgSR-NOM(aq), with a log concentration
of -13.9±0.7 M (Figure S5, Paper III).
20
Log modeled MeHg (M)
-9.5
-12.5
-11.5
-10.5
-9.5
Medium H2S(aq): 1-20 µM
Low H2S(aq): <1 µM
-10.5
y = 0.77x -2.8
R² = 0.75
-11.5
1:1
Log measured MeHg (M)
High H2S(aq): 20-60 µM
-12.5
Figure 3. Relationship between modeled and measured solubility of MeHg. The log K for
the formation of MeHgSH0(aq) was 14.6 and MeHgSR-NOM was 17.5. The pKa values of
H2S(aq) and RSH group was 7.0 and 9.0, respectively.
3.3.3. Chemical speciation model for HgII
Model fits for HgII solubility were evaluated by comparing modeled and observed
distribution of HgII between solid/adsorbed and aqueous phases (log Kd, L kg-1).
Inorganic divalent mercury, in contrast to MeHg, forms a solid phase i.e. HgS(s).
The formation of HgS(s) has strong influence on the chemical speciation of HgII,
its solubility and availability for HgII transformation processes. The log Kd is thus
considered an appropriate parameter to evaluate the HgII model fit.
Schwarzenbach and Widmer reported stability constants (log K) for the
formation of Hg(SH)20(aq) and HgS(s) of 38.6 and 37.6, respectively.96 Recently,
Drott et al. determined those stability constants and reported a log K of 36.8 for
the formation of crystalline metacinnabar (β-HgS(s)) and a log K of 39.1 for the
formation of the aqueous complex (Hg(SH)20(aq)).37 The best fit (merit-offit=0.041) was obtained with log K of 38.6 for the formation of Hg(SH) 20(aq) and
37.3 for the formation of HgS(s) as shown in Figure 4. Hence, the best fitted log K
value for the formation of HgS(s) was 0.3 units below the constant reported by
21
Schwarzenbach and Widmer96 and 0.5 units above the constant reported by Drott
et al.37 The constant for the formation of Hg(SH)20(aq) was in agreement with the
value reported by Schwarzenbach and Widmer.96 With these constants, the
average log Kd for observation (4.04±0.35) and prediction by the model
(3.97±0.68) were in good agreement. The influence of log K for the formation of
Hg(NOM-RS)2 was also illustrated by varying the log K value from 38.0 to 42.0.
The results showed that the log K of 40.0 (and pKa = 9.0) gave a better fit than
either lower or higher stability of Hg(NOM-RS)2 complex. This is in excellent
agreement with the stability constant of Hg(NOM-RS)2 previously reported by
Skyllberg.25
Although the average modeled and observed log Kd are in a good agreement, the
model showing the best merit-of-fit (Figure 4) predicted a larger variability in
HgII solubility, largely driven by the H 2S(aq) concentration, than what was
observed by measurements. One possible explanation for this could be that
colloidal Hg is passing through the filter membrane (0.22 µm) used to filter the
porewater samples. It has been shown that nanoparticulate HgS(s) is preserved
in the presence of LMM thiols and NOM-RSH(aq) by stabilizing clusters of HgS
and thus inhibit aggregation and crystallization.97-99
22
Measured HgII log Kd (L kg-1)
6
1:1
5
4
3
Low H2S(aq): <1 µM
2
Medium H2S(aq): 1-20 µM
High H2S(aq): 20-60 µM
1
1
2
3
4
II
5
6
-1
Modeled Hg log Kd (L kg )
Figure 4. Observed and modeled distribution of HgII between solid/adsorbed and
aqueous phases (log Kd, L kg-1). The log K values for the formation of Hg(SH)20(aq),
HgS(s) and Hg(NOM-RS)2 are 38.6, 37.3 and 40.0, respectively.
3.3.4. Formation of HgS(s)
HgS(s) is formed when the ion activity product [H+]/([Hg2+]×[HS-]) exceeds the
value of the stability constant for the reaction Hg 2+ +HS-= HgS(s) + H+. Hence,
free concentrations of the Hg2+ ion and inorganic sulfide together with pH control
the formation of HgS(s). Essentially, the formation of HgS(s) is controlled by two
processes: (1) the competition between dissolved aqueous HgII–sulfides
complexes and HgS(s) at high S(-II) concentration, and (2) the competition
between the complexation of Hg2+ with NOM-RSH(ads+aq) and HgS(s) at low S(II)
concentration. As illustrated in Figure 5, more than 10 mM of S (-II) is required
for Hg(SH)20(aq) to outcompete and inhibit the formation of HgS(s). In the type
of wetland soils investigated in this study, S(-II) concentrations never reach 10 mM
and the formation of HgS(s) is thus never controlled by dissolved HgII–sulfides
species. When it comes to the competition with NOM-RSH, three scenarios are
illustrated in Figure 5 based on the measured concentrations of NOM-RSH(ads)
in the wetland soils. The first scenario covers the lowest measured concentration
23
(0.26 mmol L-1), the second covers an intermediate (the average of all
concentrations = 2.6 mmol L-1) and the third comprises the highest observed
concentration (7.1 mmol L-1). At the lowest concentration of NOM-RSH(ads), the
formation of HgS(s) is initiated at 6 nM and equal amounts of HgS(s) and
Hg(NOM-RS)2(ads) are reached at 10 nM of S(-II). At the average NOM-RSH(ads)
concentration the threshold for HgS(s) formation is 0.5 µM and equal amounts of
HgS(s) and Hg(NOM-RS)2(ads) are reached at 0.9 µM S(-II). At the highest
concentration of NOM-RSH(ads) at least 4 µM of S(-II) is required to initiate the
formation of HgS(s) and equal amounts of HgS(s) and Hg(NOM-RS)2(ads) are
reached at approximately 7 µM of S(-II). This means that depending on the NOM
content in wetland soils, the threshold for HgS(s) formation may vary between 6
nM and 4 µM of S(-II).
HgS(s) (a)
HgS(s) (b)
-7.0
HgS(s) (c)
Hg(NOM-RS)2(ads) (a)
-7.5
Hg(NOM-RS)2(ads) (b)
Hg(NOM-RS)2(ads) (c)
-8.0
Hg(SH)20(aq) (a, b, c)
Log HgII species concentration (mol L-1)
-6.5
-8.5
-9
-8
-7
-6
-5
-4
-3
-2
-1
Log S(-II) concentration (M)
Figure 5. Concentration of HgII species in solid/adsorbed phase per liter of porewater, as
a function of S(-II) concentrations in boreal wetlands. The HgII concentration and pH value
were kept constant at the average measured concentrations of 82 nmol L -1 and 5.0,
respectively. The NOM-RSH(ads) concentration was varied from the low to high measured
concentration in the wetland soils: (a) low (0.26 mmol L-1), (b) intermediate (2.6 mmol L1),
and (c) high (7.1 mmol L-1). The intersections between HgS(s) and Hg(NOM-RS)2(ads)
demonstrate equal amounts of those species. Lines representing HgS(s) begin at the
threshold point for the initiation of HgS(s) precipitation. The log K values for the
formation of Hg(SH)20(aq), HgS(s) and Hg(NOM-RS)2(ads) are 38.6, 37.3 and 40.0,
respectively.
24
3.4. Factors controlling MeHg formation (Paper IV &V)
3.4.1. Determination of HgII methylation rate constant
Dissolved HgII (HgII(aq)) enriched isotope are typically used as tracers for the
determination of HgII methylation rate constant (km).36,
56,
60,
66
In the
environment, HgII is predominantly associated with NOM in the solid/adsorbed
phase and/or occurs as HgS(s). These forms of HgII are less available for the
methylation as compared to dissolved HgII(aq) tracers.48,
67
The methylation of
added dissolved HgII(aq) tracers is somewhat overestimating the methylation
capacity of the environments. In this study, two HgII tracers:
200HgII-NOM(ads)
prepared from
equilibrating
The
200HgII
were used to determine the km. The
198HgII-nitrate
200HgII-nitrate
in water while
198HgII(aq)
198HgII(aq)
200HgII-NOM(ads)
and
tracer was
was prepared by
with natural organic matter from an organic soil.68
tracer was demonstrated to be bound to RSH groups of NOM as
Hg(NOM-RS)2(ads) complex due to the strong affinity of HgII for RSH functional
groups.26, 27
Figure 6 shows an excellent correlation (R2 = 0.88) for the km determined by the
two tracers, with approximately 5 times higher methylation rate of
compared to
200HgII-NOM(ads).
between
and
as
This result is in an agreement with results from
previous studies of an estuary sediment.48,
198HgII(aq)
198HgII(aq)
200HgII-NOM(ads)
67
A strong positive correlation
tracers indicates that the
200HgII-
NOM(ads) tracer is possibly useful for the determination of ambient HgII
methylation in natural environments in which Hg(NOM-RS)2 complexes
dominate. It also shows that the 198HgII(aq) tracer gives results which can be used
to compare relative methylation rates for different sites/samples/conditions even
if the absolute rates are overestimated in comparison to methylation rates of
ambient HgII in the system.
25
Figure 6. Correlation between the methylation rate constants (km) determined for
dissolved labile 198HgII(aq) and adsorbed 200HgII-NOM(ads) tracers.
3.4.2. Effect of microbial activity on labile HgII tracer added ex situ
As mentioned in part 3.1.2, microorganisms excrete biological (Bio) LMM
thiols.61, 62, 100 If they are formed as a consequence of metabolism, concentrations
of Bio-LMM thiols can be expected to correlate positively with microbial activity.
If so, Bio-LMM thiols may be used as proxy for potential MeHg formation within
a wetland site. In the boreal wetlands (Paper IV), we observed a positively
significant correlation between total concentrations of Bio-LMM thiols and km
with R2 coefficient greater than 0.72 for samples taken at the wetlands LDN, GTN
and SKM sites (Figure 7). The fourth site: KSN had a very low concentration of
direct Bio-LMM thiols, but showed high concentrations of Glyc which was most
likely originating from both direct and indirect biological processes.101 With an
assumption that Glyc belongs to the group of Bio-LMM thiols, total LMM thiol
concentration at KSN also showed a fairly strong positive correlation with km
(Figure 7, R2= 0.41). Within the sampling site, the methylation rate in the sample
with presumably highest microbial activity is 8, 10, 30 and 70 times higher as
26
compared to the lowest one in SKM, KSN, GTN and LDN sites, respectively
(Figure 7).
SKM
0.06
R² = 0.72
0.05
Km (d-1)
KSN
0.08
0.06
R² = 0.41
0.04
0.03
0.04
0.02
0.02
0.01
0
0
0
5
10
15
20
0
Total Bio- LMM thiols (nM)
0.07
0.1
LDN
0.06
50
100
150
Total LMM thiols (nM)
GTN
R² = 0.72
0.08
R² = 0.80
Km (d-1)
0.05
0.06
0.04
0.03
0.04
0.02
0.02
0.01
0
0
0
10
20
30
Total Bio- LMM thiols (nM)
40
0
5
10
15
Total Bio-LMM thiols (nM)
Figure 7. Relationship between the methylation rate constant (km) of the
198HgII(aq)
tracer and LMM thiols concentration in the four boreal wetlands.
In the estuarine mesocosm-scale model ecosystem (Paper V), the microbial
activity (or biomass) of the pelagic system was increased following nutrient
loading to the water column, as reflected by the increase in chlorophyll α (Chl α)
and the rate of photosynthetic primary biomass production (P.P) as shown in
Figure 8a. The sediment samples were collected at different time points to
determine km. Following the increase of microbial activity, km also increased and
reached a maximum at the condition with highest microbial activity in Figure 8b.
The km value was approximately 4 times higher at the highest microbial activity
conditions102 as compared to the moderate conditions103. A significant correlation
between km and Chl α concentration (R2=0.96) was observed. After day 22 of the
27
experiment Chl α and P.P decreased, despite the maintained nutrient loading.
This may be explained by the growth rate of predatory zooplankton which
commonly increases subsequent to phytoplankton growth, generating a predator
prey‒dynamics.104
The effect of microbial activity on km in the wetlands (Paper IV) and in the
estuarine sediment (Paper V) point towards the conclusion that bacterial activity
limits the methylation rate of HgII species being a high bioavailability.
Figure 8. (a) Increase of chlorophyll α (Chl α) and the rate of photosynthetic primary
biomass production (P.P) in response to increased nutrient (N, P) addition. (b)
Methylation rate constant (km) of
199HgII(aq)
in the sediment samples as determined at
different time during the experiment.
3.4.3. Effect of microbial activity on net methylation in situ
The MeHg/HgII molar ratio is frequently used to investigate the net MeHg
formation in natural environments.48, 56, 60 In contrast to a significant correlation
between km and Bio-LMM thiols concentration in the SKM, LDN and GTN sites
(Figure 7), the MeHg/HgII molar ratios showed a weak correlation (nonsignificant, R2=0.18) to the Bio-LMM thiols concentration as shown in Figure 9.
Considering the fact that HgII is predominantly associated with thiol groups in
28
NOM in the solid/adsorbed phase and as an solid phase: HgS(s),25, 26 it is much
less available for methylation when compared to dissolved labile HgII(aq)
tracers.48, 67
Figure 9. Correlation between total biological (Bio) LMM thiol concentrations in
porewaters at the SKM, LDN and GTN wetland sites and MeHg/HgII molar ratios in the
wetland soils.
In the studied estuarine sediment (Paper V), km of the dissolved labile 199HgII(aq)
tracer increased 4 times at the maximum nutrient conditions. In contrast to this,
the increase in MeHg/HgII molar ratio of the
200HgII-NOM(ads)
tracer and
ambient HgII was not significant following increased nutrient addition (Figure
10). The different response in HgII methylation of the
and ambient
HgII,
as compared with the
HgII(aq)
200HgII-NOM(ads)
tracer
tracer, is explained by the
different availability of HgII from these three sources for methylation. The
200HgII-NOM(ads)
tracer was pre-equilibrated as an organic soil slurry in
degassed MQ water for one week prior to the experiment and200HgII is expected
to be strongly bind to RSH groups in the NOM, reducing its availability for
methylation. Ambient HgII is also expected to predominantly present as
solid/adsorbed species i.e. Hg(NOM-RS)2(ads) and HgS(s).
29
The results of MeHg formation ex situ and in situ supported the conclusion that
MeHg formation in the environment is driven by two factors: (1) the activity of
methylating bacteria, and (2) the availability of HgII for methylation. It however
remains to be clarified whether there is a breaking point at which one of the two
factors take over the control in natural environments.
Nutrient
200
HgII-NOM(ads)
Ambient Hg
150
120
0.03
90
0.02
60
0.01
Nutrient (µg L-1)
MeHg/HgII molar ratio
0.04
30
0
0.00
0
5
10
15
20
25
30
35
Time (d)
Figure 10. Responding of MeHg/HgII molar ratios in the sediment according to nutrient
loading for 200HgII-NOM(ads) and ambient Hg.
3.4.4. Net methylation and demethylation of different Hg pools
Different Hg tracers were added to the sediment at 1 cm depth below the surface
(β-201HgSsed,
200HgII-NOM
sed,
Me198Hg-NOMsed) and to the water column
(204HgIIwt and Me199Hgwt). The β-201HgSsed,
200HgII-NOM
sed
and Me198Hg-NOMsed
tracers represent pools of HgS(s) and organically complexed Hg II and MeHg in
the sediment, whereas
204HgII
wt
and Me199Hgwt represent the sum of atmospheric
deposition and terrestrial run off for HgII and MeHg. The average net
methylation as reflected by the MeHg/HgII molar ratio, was significantly different
for the different HgII tracers (ANOVA, p<0.05). The highest value was obtained
for the
204HgII ,
wt
0.041 ± 0.013, intermediate for
200HgII-NOM
sed,
0.020 ± 0.004
and lowest for β-201HgSsed tracer, 0.004 ± 0.006 (Figure 2a, Paper V). These
30
results indicate a high availability for methylation of HgII deposited in the water
phase.
For the two MeHg tracers, the net demethylation was significantly lower for the
Me198Hg-NOMsed compared to Me199Hgwt, as manifested by a higher average
MeHg/HgII molar ratio of 0.34 ± 0.07 for Me198Hg-NOMsed compared to 0.14 ±
0.04 for Me199Hgwt (Figure 2b, Paper V). This could be explained by longer and
more intense light exposed to the Me199Hgwt (in water column) compared to the
Me198Hg-NOMsed (at 1 cm sediment depth).105,
rate of
Me199Hgwt
was higher than
bioaccumulation factor of
106
Me198Hg-NOM
Me199Hgwt is
Although the demethylation
sed,
it has been shown that the
much higher than Me198Hg-NOMsed.48
To summarize, HgII in aquatic ecosystem originated from atmosphere deposition
and terrestrial run off shows a higher availability for net MeHg formation and
bioaccumulation48 than the sediment pools of HgII and MeHg. This finding
strongly points at processes and activities which contribute to an overall release
of Hg into aquatic environments, such as e.g. fossil fuel burning, gold mining,
forest clear cut, and agriculture harvesting could contribute to an overall release
of Hg into aquatic environments, thus having deteriorating effects on the aquatic
ecosystem food web.6, 107-110
3.5. Remarks for the future research
3.5.1. Determination of Hg chemical speciation on the methylating
bacteria surface proximity
Despite the extensive efforts dedicated towards the elucidation of the mechanism
of the MeHg formation over the past decade, factors and processes that control
MeHg in soils and waters are still not fully understood. Both passive35 and
active38 uptake mechanism have been suggested for the uptake of available
Hg(SH)20(aq), and specific Hg(LMM-RS)2 species by methylating bacteria.
Recently, nanoparticulate HgS(s) has also been reported available for uptake of
methylating bacteria but the mechanism has not been fully described.51 The
boreal wetlands study (Paper IV) showed that neither concentrations of
Hg(SH)20(aq) nor any specific Hg(LMM-RS)2(aq) complexes in bulk wetland
porewaters were significantly correlated to MeHg formation (km and MeHg/HgII
molar ratio). It is noteworthy that bacteria (including methylating bacteria)
31
excrete LMM thiols and/or sulfide (sulfate reducing bacteria (SRB)). 41, 42, 111 The
chemical speciation of HgII on the surrounding bacterial surface is thereby likely
very much different from bulk solution in terms of LMM thiols and/or sulfide
concentration. In a recent study on biofilm, Leclerc et al. indeed reported up to 3
orders of magnitude greater concentration of extracellular LMM thiols inside
biofilms as compared to the bulk water column.100 The aqueous environment
surrounding methylating bacteria can be expected to control the HgII uptake by
changing the chemical speciation of HgII as illustrated in Figure 11.
It is however challenging to isolate and quantify concentrations of the chemical
species at this small scale and low concentrations. The definition for the bacterial
surface vicinity is not given, this is however expected to be correlated with
bacterial dimension. Techniques such as Raman spectroscopy, Fourier transform
Infrared Spectroscopy (FTIR), transmission electron microscopy (TEM), and
synchrotron are suggested for exploring the issues.
Figure 11. Illustration for HgII chemical speciation at the near surface environment of
HgII methylating bacterium. The figure illustrates for two different types of methylating
bacteria that are iron reducing bacteria (IRB, green zone) and sulfate reducing bacteria
(SRB, yellow zone). Methylating bacteria can produce both LMM thiols and inorganic
sulfide. Based on the properties of both bacteria strains and in order to simplify the
illustration, LMM thiols (Cys as representative) and (S(-II)) are assumed to be excreted by
IRB and SRB, respectively.
32
3.5.2. Experimental approaches for determination of MeHg, and HgII
complexes to LMM thiols in pure bacteria culture
In order to improve understanding on the mechanism of MeHg formation, the
actual chemical forms of HgII and MeHgII must be experimentally determined to
elucidate the HgII uptake and the MeHg secretion processes. We are therefore
developing robust novel analytical method for determination of such complexes
based on liquid chromatography (LC) and mass spectrometry (MS) detection.
For the determination of MeHg-LMM thiol complexes, preliminary results show
that 8 different MeHg-LMM thiol complexes were detected in pure Geobacter
sulfurreducens pca culture experiment by amendment of 100 nM of HgII(aq) into
the growing medium with the bacteria optical density of 0.09-0.1 at 660 nm wave
length. The LODs for 14 investigated MeHg-LMM thiol complexes range from pM
to low nM concentration. This indicates that the method can be applied to study
the transformation of HgII in bacteria culture with amended HgII concentration
relevant to natural environments as well as the bioaccumulation of different
MeHg-LMM thiol complexes in phytoplankton.
For the determination of HgII-LMM thiol complexes, 15 different complexes have
been successfully separated and detected by LC coupled to ICPMS. Detection
limits in the sub nM range is expected for the optimized method.
33
4. Conclusions
This thesis developed new techniques for the determination of LMM thiols in the
natural environments and stability constants of relatively large number of
Hg(LMM-RS)2 complexes (15 complexes). By using the same methodology, the
relative binding strength of different LMM thiols to HgII are directly compared.
We contributes some important aspects to improve the understanding on MeHg
formation in the natural environment.

LMM thiols are frequently detected in boreal wetlands and their
concentration.
In the boreal wetlands 8 different LMM thiols (Cys, MAC, GSH, NACCys, SUC,
GluCys, 2-MPA, Glyc) were detected out of which Cys and MAC were the most
abundant with the concentration of individual LMM thiols ranging from 0.1 nM
(LOD) to 130 nM. The developed analytical method presented (as presented in
Paper I) can be applied to various type of natural waters such as freshwater,
wetland porewater, marine and lake sediment porewaters.

Thermodynamically stable for complexes between HgII and LMM thiols
This thesis also reports more precise stability constant for Hg(LMM-RS)2
complexes based on direct measurement of complexes. The stability constant for
HgII complexation with LMM thiols as expressed by the reaction Hg2+ + LMMRS- = Hg(LMM-RS)2 ranges from 34.6 for Hg(CysGly)2 to 42.1 for Hg(3-MPA)2
for 15 investigated LMM-RSH compounds. The Hg(LMM-RS)2 complexes can
therefore be classified as thermodynamically very stable. The stability constants
for 15 frequently reported LMM thiols complexed by HgII are presented in Paper
II.

Concentration of MeHg species in boreal wetlands
The research reports MeHgSR-NOM(ads) as the most abundant chemical form of
MeHg with an average concentration of 5 nmol kg-1. In the aqueous phase, the
concentration of MeHgSR-NOM(aq) dominates with an average of 10 pM,
followed by MeHgSH0(aq) of 3 pM and MeHgSR-LMM(aq) of 0.03 pM.
34

Concentration of HgII species in boreal wetlands
In the solid/adsorbed phase, the average amounts of HgS(s) and Hg(NOMRS)2(ads) are 60 and 30 nmol kg-1, respectively. In the aqueous phase, the
average concentration of Hg(SH)20(aq) is the dominant with 16 pM. Average
concentrations of HgSnSH-(aq)and Hg(NOM-RS)2(aq) are 2 pM and 0.1 pM,
respectively. Due to the very low concentration of LMM thiols in bulk porewaters,
the average log concentration of Hg(LMM-RS)2(aq) complexes are determined to
be as low as -19 M.

Conditions under which HgS(s) form in boreal wetlands
The formation of HgS(s) depends on the concentrations of HgII, NOM-RSH(ads)
and S(-II). At the average concentration of HgII and NOM-RSH(ads) in the boreal
wetlands in this study of 82 nmol L-1and 2.65 mmol L-1, respectively, HgS(s)
begins to form at an S(-II) concentration of 0.5 µM. When S(-II) concentration
exceeds 30mM, the Hg(SH)20(aq) complex constrains the formation of HgS(s)
but such high S(-II) concentration is not present in most boreal wetland
porewaters.

Factors controlling MeHg formation
MeHg formation is driven by two factors: (1) the activity of methylating bacteria,
and (2) the availability of HgII for methylation. It remains to be elucidate whether
there is a breaking point at which one of the two factors take over the control in
natural environments. The microbial activity (including methylating bacteria) is
provoked by nutrient loading, seasonal fluctuation and flooding via agricultural
activities and climate change. The availability of HgII is controlled by the
distribution of HgII between dissolved and solid/adsorbed phases and depends
heavily on the S(-II) concentration in porewater and the RSH(aq)/RSH(ads) molar
ratio.
35
5. Acknowledgements
I would take this opportunity to express my greatly indebted to my supervisor
Erik Björn. Thank you very much for offering me the opportunity to work in the
interesting projects. I am greatly appreciated for your nice and patient
supervision, I have learnt many things from your supervision. Thank you for your
“wine lessons” and “group activities”, that was unforgettable. It was remarkable
and my honor to be with you in the international mercury conferences. Thousand
thanks for joining the wetlands sampling campaign with me, for your cooking
and supporting, it was a tough time but we knew we leant somethings from that.
Ulf Skyllberg, my co-supervisor, I would also take this opportunity to show my
greatly grateful to you. “Miljon tacks” to you for being super nice supervisor, for
your thoughtful helps and explanations, I have learnt a lot from you. Thank you
for joining the mercury conference in Jeju and greatly hiking trip there.
I would deeply grateful our Vietnamese Big Father, Lars Lundmark, thank you
for open the door for many Vietnamese students to go to Umeå for studying in
chemistry. I am greatly indebted your incredible supports for the wetlands
sampling campaign, without you we could not make it. Rest in peace, Big Father.
Sofi Jonsson, I would give many thanks to you for inspiriting me to the
“mercury world”, thank you for joining me to the conferences and fantastic
drinking company. I am deeply thankful to Peter Hedlund for your immense
help with WinSGW. Many thanks to Jerker Fick, Richard Lindberg and
Marcus Östman for your help with LC-MSMS. Lars Lambertsson, Yu Song
(Rain), Tao Jiang (Tower), Wei Zhu, Gbotemi Adediran, Andreas Drott and
Aleksandra Skrobonja, my colleagues, thank you very much for your huge
supports in the Lab and scientific discussions.
I would give special thank to Phuoc Dinh for your valuable advices and being a
best friend. Many thanks to Hoa Vo, Chau Phan, Minh Anh Nguyen for being
best friends and your supports. I am grateful to Solomon Tesfalidet for your
36
encouragements, and your wife for her fantastic handmade coffee. Thanks to
Helen Genberg, Rose-Marie Kronberg, Erik Steinvall for nice fika and
drinking company. I would like to thank Ngoc Tung Pham for operating our
badminton group, and Tinh Hoang for your ardor helps. Thanks to Tung
Nguyen, Tri Tran, Trung Le, Cuong Pham, Hung Tran, Khanh Nguyen,
Minh Tuan Nguyen, and the other Vietnamese friends for your supports and
being nice company.
To my family. Con Cám ơn bố mẹ đã sinh ra con, nuôi dưỡng con, cho con cơ
hội được học tập và trưởng thành. Con tự hào và hạnh phúc được làm con của bố
mẹ, chúc bố mẹ luôn mạnh khỏe, hạnh phúc. Cám ơn vì là chị của em, chị Hiền,
cám ơn anh Tỉnh vì trở thành thành viên trong gia đình. Chúc Tiến Thành và Gia
Phúc chăm ngoan, học giỏi và trưởng thành.
To my family in law. Con rất vui vì trở thành một thành viên trong gia đình. Con
cám ơn những lời khuyên của bố mẹ. Chúc bố mẹ luôn mạnh khỏe, vui vẻ, hạnh
phúc.
Finally, I am hugely indebted to my beloved wife, Na Pham. Thank you very
much for coming together with me, for your cooking and everything. Wishing
you soon complete your tasks and dreams.
37
6. References
1.
USEPA Agency for Toxic Subtances & Disease Registry, Mercury in your environment.
2015.
2.
UNEP Sources, Emissions, Releases and Environmental Transport. UNEP Chemicals Branch,
Geneva, Switzerland. 2013.
3.
USEPA EPA Mercury Study Report to Congress; Office of Air Quality and Standards and
Office of Research and Developement Washington (DC), 1997.
4.
Bjorn, E.; Larsson, T.; Lambertsson, L.; Skyllberg, U.; Frech, W., Recent advances in
mercury speciation analysis with focus on spectrometric methods and enriched stable
isotope applications. Ambio 2007, 36, (6), 443-451.
5.
Selin, N. E., Global Biogeochemical Cycling of Mercury: A Review. Annual Review of
Environment and Resources 2009, 34, 43-63.
6.
Malm, O.; Pfeiffer, W. C.; Souza, C. M.; Reuther, R., Mercury pollution due to gold mining
in the Madeira River basin, Brazil. Ambio 1990, 19, (1), 11-15.
7.
Energy,
U.
S.
D.
o.
Mercury
Emission
Control
R&D.
http://www.fossil.energy.gov/programs/powersystems/pollutioncontrols/overview_merc
urycontrols.html (2 november 2011),
8.
Lamborg, C. H.; Hammerschmidt, C. R.; Bowman, K. L.; Swarr, G. J.; Munson, K. M.;
Ohnemus, D. C.; Lam, P. J.; Heimbürger, L.-E.; Rijkenberg, M. J.; Saito, M. A., A global
ocean inventory of anthropogenic mercury based on water column measurements.
Nature 2014, 512, (7512), 65-68.
9.
UNEP Chemical Branch, D. The Global Atmosperic Mercury Assessment: Sources, Emissions
and Transport; Geneva, 2008.
10.
Pacyna, E. G.; Pacyna, J. M.; Sundseth, K.; Munthe, J.; Kindbom, K.; Wilson, S.;
Steenhuisen, F.; Maxson, P., Global emission of mercury to the atmosphere from
anthropogenic sources in 2005 and projections to 2020. Atmos. Environ. 2010, 44, (20),
2487-2499.
11.
Kurland, T.; Faro, S. N.; Siedler, H., Minamata disease. The outbreak of a neurologic
disorder in Minamata, Japan, and its relationship to the ingestion of seafood
contaminated by mercuric compounds. World neurology 1960, 1, (5), 370-95.
12.
Harada, M., Minamata disease: methylmercury poisoning in Japan caused by
environmental pollution. Crit. Rev. Toxicol. 1995, 25, (1), 1-24.
13.
Agency, U. S. E. P., National Listing of Fish Advisories" (EPA Fact Sheet EPA-823-F-09-007).
2008.
14.
Jiang, G. B.; Shi, J. B.; Feng, X. B., Mercury pollution in China. Environmental Science &
Technology 2006, 40, (12), 3672-3678.
15.
Hakanson, L.; Nilsson, A.; Andersson, T., MERCURY IN FISH IN SWEDISH LAKES. Environ.
Pollut. 1988, 49, (2), 145-162.
16.
Zhang, H.; Feng, X.; Larssen, T.; Qiu, G.; Vogt, R. D., In inland China, rice, rather than fish, is
the major pathway for methylmercury exposure. Environ. Health Perspect. 2010, 118, (9),
1183.
38
17.
Union, E., Directive 2008/105/EC of the European parliament and of the council. Official
journal of the European Union 2008.
18.
Lindqvist, O.; Johansson, K.; Aastrup, M.; Andersson, A.; Bringmark, L.; Hovsenius, G.;
Hakanson, L.; Iverfeldt, A.; Meili, M.; Timm, B., MERCURY IN THE SWEDISH ENVIRONMENT
- RECENT RESEARCH ON CAUSES, CONSEQUENCES AND CORRECTIVE METHODS. Water Air
and Soil Pollution 1991, 55, (1-2), R11-&.
19.
Chemicals, U., Global mercury assessment. United Nations Environmental Programme.
Chemicals. Inter-Organization Programme for the Sound Managements of Chemicals.
Geneva, Switzerland, 258pp 2002.
20.
Bellanger, M.; Pichery, C.; Aerts, D.; Berglund, M.; Castaño, A.; Cejchanová, M.; Crettaz, P.;
Davidson, F.; Esteban, M.; Fischer, M. E., Economic benefits of methylmercury exposure
control in Europe: monetary value of neurotoxicity prevention. Environ Health 2013, 12,
(3).
21.
Clarkson, T. W.; Magos, L., The toxicology of mercury and its chemical compounds. Crit.
Rev. Toxicol. 2006, 36, (8), 609-662.
22.
Qiu, G.; Feng, X.; Li, P.; Wang, S.; Li, G.; Shang, L.; Fu, X., Methylmercury accumulation in
rice (Oryza sativa L.) grown at abandoned mercury mines in Guizhou, China. J. Agric. Food.
Chem. 2008, 56, (7), 2465-2468.
23.
Templeton, D. M.; Ariese, F.; Cornelis, R.; Danielsson, L. G.; Muntau, H.; Van Leeuwen, H.
P.; Lobinski, R., Guidelines for terms related to chemical speciation and fractionation of
elements. Definitions, structural aspects, and methodological approaches (IUPAC
Recommendations 2000). Pure Appl. Chem. 2000, 72, (8), 1453-1470.
24.
Jonsson, S., Unraveling the importance of solid and adsorbed phase mercury speciation
for methylmercury formation, evasion and bioaccumulation. 2013.
25.
Skyllberg, U., Competition among thiols and inorganic sulfides and polysulfides for Hg and
MeHg in wetland soils and sediments under suboxic conditions: Illumination of
controversies and implications for MeHg net production. Journal of Geophysical ResearchBiogeosciences 2008, 113, G00C03.
26.
Skyllberg, U., Chemical speciation of mercury in soil and sediment. Environmental
chemistry and toxicology of mercury 2012, 219-258.
27.
Skyllberg, U.; Bloom, P. R.; Qian, J.; Lin, C. M.; Bleam, W. F., Complexation of mercury(II) in
soil organic matter: EXAFS evidence for linear two-coordination with reduced sulfur
groups. Environmental Science & Technology 2006, 40, (13), 4174-4180.
28.
Dyrssen, D.; Wedborg, M., The sulphur-mercury (II) system in natural waters. Water Air &
Soil Pollution 1991, 56, (1), 507-519.
29.
Mah, V.; Jalilehvand, F., Glutathione complex formation with mercury (II) in aqueous
solution at physiological pH. Chem. Res. Toxicol. 2010, 23, (11), 1815-1823.
30.
Koszegi-Szalai, H.; Paal, T., Equilibrium studies of mercury (II) complexes with
penicillamine. Talanta 1999, 48, (2), 393-402.
31.
Tossell, J.; Vaughan, D., Relationships between valence orbital binding energies and crystal
structures in compounds of copper, silver, gold, zinc, cadmium, and mercury. Inorg. Chem.
1981, 20, (10), 3333-3340.
39
32.
Reid, R. S.; Rabenstein, D. L., Nuclear magnetic resonance studies of the solution
chemistry of metal complexes. XVII. Formation constants for the complexation of
methylmercury by sulfhydryl-containing amino acids and related molecules. Can. J. Chem.
1981, 59, (10), 1505-1514.
33.
Karlsson, T.; Skyllberg, U., Bonding of ppb levels of methyl mercury to reduced sulfur
groups in soil organic matter. Environmental Science & Technology 2003, 37, (21), 49124918.
34.
Skyllberg, U.; Xia, K.; Bloom, P. R.; Nater, E. A.; Bleam, W. F., Binding of mercury (II) to
reduced sulfur in soil organic matter along upland-peat soil transects. Journal of
environmental quality 2000, 29, (3), 855-865.
35.
Benoit, J. M.; Gilmour, C. C.; Mason, R. P.; Heyes, A., Sulfide controls on mercury
speciation and bioavailability to methylating bacteria in sediment pore waters.
Environmental Science & Technology 1999, 33, (6), 951-957.
36.
Drott, A.; Lambertsson, L.; Bjorn, E.; Skyllberg, U., Importance of dissolved neutral
mercury sulfides for methyl mercury production in contaminated sediments.
Environmental Science & Technology 2007, 41, (7), 2270-2276.
37.
Drott, A.; Björn, E.; Bouchet, S.; Skyllberg, U., Refining Thermodynamic Constants for
Mercury(II)-Sulfides in Equilibrium with Metacinnabar at Sub-Micromolar Aqueous Sulfide
Concentrations. Environmental Science & Technology 2013, 47, (9), 4197-4203.
38.
Schaefer, J. K.; Rocks, S. S.; Zheng, W.; Liang, L. Y.; Gu, B. H.; Morel, F. M. M., Active
transport, substrate specificity, and methylation of Hg(II) in anaerobic bacteria.
Proceedings of the National Academy of Sciences of the United States of America 2011,
108, (21), 8714-8719.
39.
Liem-Nguyen, V.; Bouchet, S.; Bjö rn, E., Determination of Sub-Nanomolar Levels of Low
Molecular Mass Thiols in Natural Waters by Liquid Chromatography Tandem Mass
Spectrometry after Derivatization with p-(Hydroxymercuri) Benzoate and Online
Preconcentration. Anal. Chem. 2014, 87, (2), 1089-1096.
40.
Schaefer, J. K.; Morel, F. M. M., High methylation rates of mercury bound to cysteine by
Geobacter sulfurreducens. Nature Geoscience 2009, 2, (2), 123-126.
41.
Gilmour, C. C.; Henry, E. A.; Mitchell, R., Sulfate Stimulation of Mercury Methylation in
Fresh-Water Sediments. Environmental Science & Technology 1992, 26, (11), 2281-2287.
42.
Compeau, G. C.; Bartha, R., SULFATE-REDUCING BACTERIA - PRINCIPAL METHYLATORS OF
MERCURY IN ANOXIC ESTUARINE SEDIMENT. Applied and Environmental Microbiology
1985, 50, (2), 498-502.
43.
Kerin, E. J.; Gilmour, C. C.; Roden, E.; Suzuki, M. T.; Coates, J. D.; Mason, R. P., Mercury
methylation by dissimilatory iron-reducing bacteria. Applied and Environmental
Microbiology 2006, 72, (12), 7919-7921.
44.
Fleming, E. J.; Mack, E. E.; Green, P. G.; Nelson, D. C., Mercury methylation from
unexpected sources: Molybdate-inhibited freshwater sediments and an iron-reducing
bacterium. Applied and Environmental Microbiology 2006, 72, (1), 457-464.
45.
Hamelin, S.; Amyot, M.; Barkay, T.; Wang, Y.; Planas, D., Methanogens: principal
methylators of mercury in lake periphyton. Environmental science & technology 2011, 45,
(18), 7693-7700.
40
46.
Gilmour, C. C.; Podar, M.; Bullock, A. L.; Graham, A. M.; Brown, S. D.; Somenahally, A. C.;
Johs, A.; Hurt Jr, R. A.; Bailey, K. L.; Elias, D. A., Mercury methylation by novel
microorganisms from new environments. Environmental science & technology 2013, 47,
(20), 11810-11820.
47.
Benoit, J. M.; Gilmour, C. C.; Heyes, A.; Mason, R. P.; Miller, C. L., Geochemical and
biological controls over methylmercury production and degradation in aquatic ecosystems.
2003; p 262-297.
48.
Jonsson, S.; Skyllberg, U.; Nilsson, M. B.; Lundberg, E.; Andersson, A.; Björn, E.,
Differentiated availability of geochemical mercury pools controls methylmercury levels in
estuarine sediment and biota. Nature communications 2014, 5.
49.
Mehrotra, A. S.; Sedlak, D. L., Decrease in net mercury methylation rates following iron
amendment to anoxic wetland sediment slurries. Environmental science & technology
2005, 39, (8), 2564-2570.
50.
Jeong, H. Y.; Klaue, B.; Blum, J. D.; Hayes, K. F., Sorption of mercuric ion by synthetic
manocrystalline mackinawite (FeS). Environmental Science & Technology 2007, 41, (22),
7699-7705.
51.
Hsu-Kim, H.; Kucharzyk, K. H.; Zhang, T.; Deshusses, M. A., Mechanisms regulating mercury
bioavailability for methylating microorganisms in the aquatic environment: a critical
review. Environmental science & technology 2013, 47, (6), 2441-2456.
52.
Zhang, J.; Wang, F.; House, J. D.; Page, B., Thiols in wetland interstitial waters and their
role in mercury and methylmercury speciation. Limnology and oceanography 2004, 49,
(6), 2276-2286.
53.
Dryden, C. L.; Gordon, A. S.; Donat, J. R., Seasonal survey of copper-complexing ligands
and thiol compounds in a heavily utilized, urban estuary: Elizabeth River, Virginia. Mar.
Chem. 2007, 103, (3), 276-288.
54.
MacCrehan, W.; Shea, D.; Vairavamurthy, M.; Schoonen, M.; Eglinton, T.; Luther III, G.;
Manowitz, B. In Temporal relationship of thiols to inorganic sulfur compounds in anoxic
Chesapeake Bay sediment porewater, ACS Symp. Ser., 1995; Washington, DC: American
Chemical Society,[1974]-: 1995; pp 294-310.
55.
Cline, J. D., SPECTROPHOTOMETRIC DETERMINATION OF HYDROGEN SULFIDE IN NATURAL
WATERS. Limnology and Oceanography 1969, 14, (3), 454-&.
56.
Tjerngren, I.; Karlsson, T.; Bjorn, E.; Skyllberg, U., Potential Hg methylation and MeHg
demethylation rates related to the nutrient status of different boreal wetlands.
Biogeochemistry 2012, 108, (1-3), 335-350.
57.
Zheng, Y.; Anderson, R. F.; van Geen, A.; Kuwabara, J., Authigenic molybdenum formation
in marine sediments: a link to pore water sulfide in the Santa Barbara Basin. Geochim.
Cosmochim. Acta 2000, 64, (24), 4165-4178.
58.
Tjerngren, I.; Meili, M.; Björn, E.; Skyllberg, U., Eight Boreal Wetlands as Sources and Sinks
for Methyl Mercury in Relation to Soil Acidity, C/N Ratio, and Small-Scale Flooding.
Environmental Science & Technology 2012, 46, (15), 8052-8060.
59.
St. Louis, V. L.; Rudd, J. W.; Kelly, C. A.; Beaty, K. G.; Bloom, N. S.; Flett, R. J., Importance of
wetlands as sources of methyl mercury to boreal forest ecosystems. Canadian Journal of
fisheries and aquatic sciences 1994, 51, (5), 1065-1076.
41
60.
Hammerschmidt, C. R.; Fitzgerald, W. F., Geochemical controls on the production and
distribution of methylmercury in near-shore marine sediments. Environmental Science &
Technology 2004, 38, (5), 1487-1495.
61.
Rao, Y.; Xiang, B.; Bramanti, E.; D’Ulivo, A.; Mester, Z., Determination of thiols in yeast by
HPLC coupled with LTQ-orbitrap mass spectrometry after derivatization with p(Hydroxymercuri) benzoate. J. Agric. Food. Chem. 2010, 58, (3), 1462-1468.
62.
Bakirdere, S.; Bramanti, E.; D’ulivo, A.; Ataman, O. Y.; Mester, Z., Speciation and
determination of thiols in biological samples using high performance liquid
chromatography–inductively coupled plasma-mass spectrometry and high performance
liquid chromatography–Orbitrap MS. Anal. Chim. Acta 2010, 680, (1), 41-47.
63.
Qian, J.; Skyllberg, U.; Frech, W.; Bleam, W. F.; Bloom, P. R.; Petit, P. E., Bonding of methyl
mercury to reduced sulfur groups in soil and stream organic matter as determined by Xray absorption spectroscopy and binding affinity studies. Geochim. Cosmochim. Acta 2002,
66, (22), 3873-3885.
64.
Burton, E. D.; Bush, R. T.; Johnston, S. G.; Sullivan, L. A.; Keene, A. F., Sulfur
biogeochemical cycling and novel Fe–S mineralization pathways in a tidally re-flooded
wetland. Geochim. Cosmochim. Acta 2011, 75, (12), 3434-3451.
65.
Lambertsson, L.; Lundberg, E.; Nilsson, M.; Frech, W., Applications of enriched stable
isotope tracers in combination with isotope dilution GC-ICP-MS to study mercury species
transformation in sea sediments during in situ ethylation and determination. J. Anal. At.
Spectrom. 2001, 16, (11), 1296-1301.
66.
Hintelmann, H.; Keppel-Jones, K.; Evans, R. D., Constants of mercury methylation and
demethylation rates in sediments and comparison of tracer and ambient mercury
availability. Environ. Toxicol. Chem. 2000, 19, (9), 2204-2211.
67.
Jonsson, S.; Skyllberg, U.; Nilsson, M. B.; Westlund, P.-O.; Shchukarev, A.; Lundberg, E.;
Björn, E., Mercury Methylation Rates for Geochemically Relevant HgII Species in
Sediments. Environmental Science & Technology 2012, 46, (21), 11653-11659.
68.
Skyllberg, U.; Drott, A., Competition between Disordered Iron Sulfide and Natural Organic
Matter Associated Thiols for Mercury(II)-An EXAFS Study. Environmental Science &
Technology 2010, 44, (4), 1254-1259.
69.
Martell, A.; Smith, R.; Motekaitis, R., National Institute of Standard and Technology, NIST.
Critically selected stability constants of metal complexes. PC-based Database,
Gaithersburg, MD 2004, 20899.
70.
Snell, J. P.; Stewart, I. I.; Sturgeon, R. E.; Frech, W., Species specific isotope dilution
calibration for determination of mercury species by gas chromatography coupled to
inductively coupled plasma- or furnace atomisation plasma ionisation-mass spectrometry.
J. Anal. At. Spectrom. 2000, 15, (12), 1540-1545.
71.
Hu, H.; Mylon, S. E.; Benoit, G., Distribution of the thiols glutathione and
3‐mercaptopropionic acid in Connecticut lakes. Limnology and oceanography 2006, 51,
(6), 2763-2774.
72.
Seiwert, B.; Karst, U., Simultaneous LC/MS/MS determination of thiols and disulfides in
urine samples based on differential labeling with ferrocene-based maleimides. Anal.
Chem. 2007, 79, (18), 7131-7138.
42
73.
Vairavamurthy, A.; Mopper, K., Field method for determination of traces of thiols in
natural waters. Anal. Chim. Acta 1990, 236, 363-370.
74.
Kuśmierek, K.; Chwatko, G.; Głowacki, R.; Kubalczyk, P.; Bald, E., Ultraviolet derivatization
of low-molecular-mass thiols for high performance liquid chromatography and capillary
electrophoresis analysis. J. Chromatogr. B 2011, 879, (17), 1290-1307.
75.
Toyo’oka, T., Recent advances in separation and detection methods for thiol compounds
in biological samples. J. Chromatogr. B 2009, 877, (28), 3318-3330.
76.
Tang, D.; Wen, L.-S.; Santschi, P. H., Analysis of biogenic thiols in natural water samples by
high-performance liquid chromatographic separation and fluorescence detection with
ammonium 7-fluorobenzo-2-oxa-1, 3-diazole-4-sulfonate (SBD-F). Anal. Chim. Acta 2000,
408, (1), 299-307.
77.
McMenamin, M. E.; Himmelfarb, J.; Nolin, T. D., Simultaneous analysis of multiple
aminothiols in human plasma by high performance liquid chromatography with
fluorescence detection. J. Chromatogr. B 2009, 877, (28), 3274-3281.
78.
Isokawa, M.; Funatsu, T.; Tsunoda, M., Fast and simultaneous analysis of biothiols by highperformance liquid chromatography with fluorescence detection under hydrophilic
interaction chromatography conditions. Analyst 2013, 138, (13), 3802-3808.
79.
Dickinson, D. A.; Forman, H. J., Glutathione in defense and signaling. Ann. N.Y. Acad. Sci.
2002, 973, (1), 488-504.
80.
Gulati, P.; Klöhn, P. C.; Krug, H.; Göttlicher, M.; Markova, B.; Böhmer, F. D.; Herrlich, P.,
Redox regulation in mammalian signal transduction. IUBMB life 2001, 52, (1), 25-28.
81.
Moriarty-Craige, S. E.; Jones, D. P., Extracellular thiols and thiol/disulfide redox in
metabolism. Annu. Rev. Nutr. 2004, 24, 481-509.
82.
Casas, J.; Jones, M. M., Mercury (II) complexes with sulfhydryl containing chelating agents:
stability constant inconsistencies and their resolution. J. Inorg. Nucl. Chem. 1980, 42, (1),
99-102.
83.
Basinger, M. A.; Casas, J.; Jones, M. M.; Weaver, A. D.; Weinstein, N. H., Structural
requirements for Hg (II) antidotes. J. Inorg. Nucl. Chem. 1981, 43, (6), 1419-1425.
84.
Starý, J.; Kratzer, K., Radiometric determination of stability constants of mercury species
complexes with L-cysteine. J. Radioanal. Nucl. Chem. 1988, 126, (1), 69-75.
85.
Cardiano, P.; Cucinotta, D.; Foti, C.; Giuffre, O.; Sammartano, S., Potentiometric,
Calorimetric, and 1H NMR Investigation on Hg2+-Mercaptocarboxylate Interaction in
Aqueous Solution. Journal of Chemical & Engineering Data 2011, 56, (5), 1995-2004.
86.
Cardiano, P.; Falcone, G.; Foti, C.; Sammartano, S., Sequestration of Hg2+ by some
biologically important thiols. Journal of Chemical & Engineering Data 2011, 56, (12), 47414750.
87.
Jalilehvand, F.; Leung, B. O.; Izadifard, M.; Damian, E., Mercury (II) cysteine complexes in
alkaline aqueous solution. Inorg. Chem. 2006, 45, (1), 66-73.
88.
Leung, B. O.; Jalilehvand, F.; Mah, V., Mercury (II) penicillamine complex formation in
alkaline aqueous solution. Dalton Transactions 2007, (41), 4666-4674.
89.
Liu, G.; Cai, Y.; O'Driscoll, N., Environmental chemistry and toxicology of mercury. John
Wiley & Sons: 2011.
43
90.
Ugur, I.; Marion, A.; Parant, S.; Jensen, J. H.; Monard, G., Rationalization of the p K a
Values of Alcohols and Thiols Using Atomic Charge Descriptors and Its Application to the
Prediction of Amino Acid p K a’s. Journal of chemical information and modeling 2014, 54,
(8), 2200-2213.
91.
Lambertsson, L.; Nilsson, M., Organic material: The primary control on mercury
methylation and ambient methyl mercury concentrations in estuarine sediments.
Environmental Science & Technology 2006, 40, (6), 1822-1829.
92.
Morel, F. M. M.; Kraepiel, A. M. L.; Amyot, M., The chemical cycle and bioaccumulation of
mercury. Annual Review of Ecology and Systematics 1998, 29, 543-566.
93.
Xia, K.; Skyllberg, U.; Bleam, W.; Bloom, P.; Nater, E.; Helmke, P., X-ray absorption
spectroscopic evidence for the complexation of Hg (II) by reduced sulfur in soil humic
substances. Environmental science & technology 1999, 33, (2), 257-261.
94.
Hesterberg, D.; Chou, J. W.; Hutchison, K. J.; Sayers, D. E., Bonding of Hg (II) to reduced
organic sulfur in humic acid as affected by S/Hg ratio. Environmental science & technology
2001, 35, (13), 2741-2745.
95.
Yoon, S.-J.; Diener, L. M.; Bloom, P. R.; Nater, E. A.; Bleam, W. F., X-ray absorption studies
of CH 3 Hg+-binding sites in humic substances. Geochim. Cosmochim. Acta 2005, 69, (5),
1111-1121.
96.
Schwarzenbach, G.; Widmer, M., Die löslichkeit von metallsulfiden I. schwarzes
quecksilbersulfid. Helv. Chim. Acta 1963, 46, (7), 2613-2628.
97.
Deonarine, A.; Hsu-Kim, H., Precipitation of mercuric sulfide nanoparticles in NOMcontaining water: Implications for the natural environment. Environmental science &
technology 2009, 43, (7), 2368-2373.
98.
Slowey, A. J., Rate of formation and dissolution of mercury sulfide nanoparticles: The dual
role of natural organic matter. Geochim. Cosmochim. Acta 2010, 74, (16), 4693-4708.
99.
Pham, A. L.-T.; Morris, A.; Zhang, T.; Ticknor, J.; Levard, C.; Hsu-Kim, H., Precipitation of
nanoscale mercuric sulfides in the presence of natural organic matter: Structural
properties, aggregation, and biotransformation. Geochim. Cosmochim. Acta 2014, 133,
204-215.
100.
Leclerc, M.; Planas, D.; Amyot, M., Relationship between extracellular low-molecularweight thiols and mercury species in natural lake periphytic biofilms. Environmental
science & technology 2015, 49, (13), 7709-7716.
101.
Shea, D.; MacCrehan, W. A., Determination of hydrophilic thiols in sediment porewater
using ion-pair liquid chromatography coupled to electrochemical detection. Anal. Chem.
1988, 60, (14), 1449-1454.
102.
Andersson, A.; Hajdu, S.; Haecky, P.; Kuparinen, J.; Wikner, J., Succession and growth
limitation of phytoplankton in the Gulf of Bothnia (Baltic Sea). Mar. Biol. 1996, 126, (4),
791-801.
103.
Andersson, A.; Haecky, P.; Hagström, Å., Effect of temperature and light on the growth of
micro-nano-and pico-plankton: impact on algal succession. Mar. Biol. 1994, 120, (4), 511520.
104.
Elmgren, R., Trophic dynamics in the enclosed, brackish Baltic Sea. Rapp Pv Réun. Cons.
Int. Explor. Mer 1984, 183, 152-169.
44
105.
Lehnherr, I.; St Louis, V. L.; Hintelmann, H.; Kirk, J. L., Methylation of inorganic mercury in
polar marine waters. Nature geoscience 2011, 4, (5), 298-302.
106.
Hammerschmidt, C. R.; Fitzgerald, W. F., Photodecomposition of methylmercury in an
arctic Alaskan lake. Environmental science & technology 2006, 40, (4), 1212-1216.
107.
Rytuba, J. J., Mercury from mineral deposits and potential environmental impact. Environ.
Geol. 2003, 43, (3), 326-338.
108.
Malm, O., Gold mining as a source of mercury exposure in the Brazilian Amazon. Environ.
Res. 1998, 77, (2), 73-78.
109.
Porvari, P.; Verta, M.; Munthe, J.; Haapanen, M., Forestry practices increase mercury and
methyl mercury output from boreal forest catchments. Environmental Science &
Technology 2003, 37, (11), 2389-2393.
110.
Kronberg, R.-M., The boreal journey of methyl mercury. 2014.
111.
Mangal, V.; Guéguen, C., Examining concentrations and molecular weights of thiols in
microorganism cultures and in Churchill River (Manitoba) using a fluorescent-labeling
method coupled to asymmetrical flow field-flow fractionation. Analytical and bioanalytical
chemistry 2015, 407, (15), 4305-4313.
45