Mercury species transformations in marine and
biological systems studied by isotope dilution
mass spectrometry and stable isotope tracers
by
Lars Lambertsson
b
AKADEMISK AVHANDLING
Som med vederbörligt tillstånd av Rektorsämbetet vid Umeå Universitet för erhållande av
filosofie doktorsexamen framlägges till offentlig granskning vid Kemiska Instutitionen, Sal
KB3A9 (Lilla hörsalen) KBC-huset, fredagen den 1 April, Klockan 10.00.
Fakultetsopponent: Dr. John Munthe, IVL Svenska Miljöinstitutet AB, Göteborg
Titel:
Mercury species transformations in marine and biological systems studied
by isotope dilution mass spectrometry and stable isotope tracers
Author:
Lars Lambertsson
Department of Chemistry, Analytical Chemistry, Umeå University,
S-90187 Umeå, Sweden
Abstract:
This thesis focuses on the implementation of species-specific isotope dilution
(SSID) methodology and stable isotope tracers to determine mercury species
occurrence and transformation processes in-situ and during sample treatment.
Isotope enriched tracers of methyl-, ethyl- and inorganic mercury were
synthesised and applied in different combinations to marine and biological
samples. Experimental results were obtained using gas chromatographyinductively coupled plasma-mass spectrometry (GC-ICP-MS).
Mercury methylation and methylmercury demethylation processes in
surface sediments were studied in the brackish Öre River estuary, Bothnian
Bay. Uni- and multivariate data evaluation identified the organic material
content and mercury methylation potential in the sediments as important
factors controlling incipient methylmercury levels. Mercury species
distribution in mice treated with the pharmaceutical preservative Thimerosal
(ethylmercurithiosalicylate) was studied. The ethylmercury moiety of
Thimerosal was observed to rapidly convert to inorganic mercury in the mice
during the treatment period as well as during sample treatment, hence
necessitating SSID methodology for accurate ethylmercury determinations in
biological samples.
To facilitate the introduction of SSID as a routine quantitative method
in mercury speciation, a methylmercury isotopic certified reference material
(ICRM) was produced. Prior to certification, the stability of the material was
examined in conventional and isochronous stability studies spanning 12
months, which permitted uncertainty estimation of the methylmercury amount
content for two years of shelf-life.
Finally, a field-adapted SSID method for methylmercury
determinations in natural water samples was developed. The proposed
analytical protocol significantly simplified sample storage- and treatment
procedures without sacrifices in analytical accuracy.
Keywords:
Mercury, methylmercury, ethylmercury, Thimerosal, speciation, methylation,
demethylation, brackish water sediment, natural waters, GC-ICP-MS, speciesspecific isotope dilution, isotope enriched tracers
ISBN:
91-7305-838-6
© Copyright, Lars Lambertsson, 2005. All rights reserved
Printed by Solfjädern Offset AB, Umeå
ii
This thesis includes the following papers, which are referred to in the text by Roman numerals
I
L. Lambertsson, E. Lundberg, M. Nilsson and W. Frech
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
Journal of Analytical Atomic Spectrometry, 2001, 16, 1296.
II
L. Lambertsson and M. Nilsson
Organic material as the primary control on net mercury methylation in estuarine
sediments
Environmental Science and Technology, submitted.
III
J. Qvarnström, L. Lambertsson, S. Havarinasab, P. Hultman and W. Frech
Determination of methylmercury, ethylmercury and inorganic mercury in mouse
tissues, following administration of Thimerosal, by species-specific isotope dilution
GC-inductively coupled plasma MS
Analytical Chemistry, 2003, 75, 4120.
IV
J. P. Snell, C. R. Quetél, L. Lambertsson and J. Qvarnström
Preparation and certification of ERM-AE670, a 202Hg enriched methylmercury
isotopic reference material
Journal of Analytical Atomic Spectrometry, 2004, 19, 1315.
V
L. Lambertsson and E. Björn
Validation of a simplified field-adapted procedure for routine determinations of
methylmercury at trace levels in natural water samples using species-specific isotope
dilution mass spectrometry
Analytical & Bioanalytical Chemistry, 2004, 380, 871.
Paper I and IV are reprinted from Journal of Analytical Atomic Spectrometry with permission from the Royal
Society of Chemistry (copyright 2001 and 2004, respectively). Paper III is reprinted from Analytical Chemistry
with permission from the American Chemical Society (copyright 2003) and paper V is reprinted from Analytical
& Bioanalytical Chemistry with permission from Springer Verlag (copyright 2004).
iii
Also by the author
S. Havarinasab, L. Lambertsson, J. Qvarnström and P. Hultman
Dose-response study of thimerosal-induced murine systemic autoimmunity
Toxicology and Applied Pharmacology, 2004, 194, 169.
iv
Table of contents
1. Introduction
1
2. Mercury
1
2
2
3
3
3
6
7
2.1
Toxicology
2.1.1 Elemental mercury
2.1.2 Inorganic mercury salts
2.1.3 Organic mercury
2.2
The mercury cycle
2.2.1 Mercury methylation
2.2.2 Methylmercury demethylation
3. Analytical techniques and methodology for mercury speciation
3.1
3.2
Sample storage, preservation and preconcentration
Mercury species separation
3.2.1 Gas Chromatography
3.3
Mercury species detection
3.3.1 ICP-MS
3.3.2 GC-ICP-MS
3.4
Isotope dilution mass spectrometry (IDMS)
3.4.1 Isotopomer synthesis
3.5
Analytical methods for monitoring mercury species
transformations in the environment
4. Results: Summary of published papers and manuscripts
4.1
4.2
4.3
Mercury methylation and methylmercury demethylation
in brackish water sediments: Improved methodology and
geochemical controls, papers I and II
Mercury species transformations in Thimerosal treated
mice monitored with multi labelled SSID, paper III
Towards a wider implementation of SSID methodology
in mercury speciation, papers IV and V
8
9
10
10
12
12
13
14
16
17
20
20
23
26
5. Conclusions and prospects
27
6. Acknowledgements
29
7. References
30
v
1.
Introduction
In nature, an element may exist in different chemical forms (species) that are characterised by
oxidation state and/or molecular structure. The toxicity and bioavailability of some elements
is often directly related to the chemical form, for example trivalent Chromium (Cr3+) is
essential for glucose metabolism in animals whereas hexavalent Chromium (Cr6+) is
carcinogenic. Many other elements, such as mercury, tin, arsenic, lead and selenium, display
similar “inter-species” toxicity differences. Obviously, it is of concern to have available
sensitive and accurate analytical methods for qualitative and quantitative element speciation
in any given sample. During the last 40 years there has been a steadily increasing interest in
the field of speciation science, which is the commonly used name for this branch of analytical
chemistry. The term speciation was recently defined by IUPAC (International Union for Pure
and Applied Chemistry) as “Determination of the exact chemical form or compound in which
an element occurs in a sample, for instance determination of whether arsenic occurs in the
form of trivalent or pentavalent ions or as part of an organic molecule, and the quantitative
distribution of the different chemical forms that may coexist” [1]. Often, establishment of the
total concentrations of many elements would not provide sufficient information in risk
assessment regarding for instance food and environmental safety. In fact, present European
legislation requires speciation of mercury, cadmium, lead, tin and nickel in certain
environmental samples and biota [2].
This thesis describes development of analytical methodology based on speciesspecific isotope dilution mass spectrometry and isotope enriched tracers for determination of
mercury species and their transformation processes in marine and biological matrices.
2.
Mercury
Mercury (Hg) is a natural constituent element of the earth crust. It exists in three oxidation
states: elemental mercury (Hg0), mercurous mercury (Hg+) and mercuric mercury (Hg2+). At
room temperature, elemental mercury (Hg0) is a silvery liquid metal, hence the latin name
“hydrargyrum” meaning water-silver. Given a high vapour pressure (0.17 Pa at 25 ° C),
however, Hg0 is easily emitted to the atmosphere and is rarely encountered as the liquid metal
in nature. A majority of the immobilised or “buried” mercury found in the earth crust are
inorganic mercuric compounds (mercury salts), e.g. mercuric oxide (HgO), mercuric chloride
(HgCl2) and in particular mercuric sulphide (HgS) or cinnabar. Cinnabar is mined to produce
elemental mercury for a range of industrial applications, for instance as a catalyst in chemical
processes and in small-scale gold mining operations. Elemental mercury is also used in
various products such as electrical switches, thermometers and medical equipment, batteries
and copper/silver amalgams tooth filling material to name a few. The recorded annual global
primary production of mercury in year 2000 was 1800 metric tonnes and in addition an
estimated 700-900 metric tonnes of recycled mercury has been marketed yearly since the mid
90’s [3].
Mercuric mercury (Hg2+), hereafter referred to as inorganic mercury, may also form
mono- or di-substituted organometallic compounds through covalent bonding to phenyl or
short-chained alkyl substituents, yielding for example dimethylmercury ((CH3)2Hg),
ethylmercury (C2H5Hg+), phenylmercury (C6H5Hg+) and, of particular environmental
concern, methylmercury (CH3Hg+).
The first documented human activity involving organic mercury compounds dates back to
the 1860’s, when methylmercury compounds were first synthesised. Apparently, two of the
1
scientists involved in the experiment died from unsafe exposure to the products. As a result,
the scientific community avoided organic mercury compounds until the beginning of the
twentieth century when it was found that methyl- and other short-chained alkyl- or
phenylmercury derivatives had excellent anti-fungal properties [4]. Organic mercury
compounds were from then on used in numerous applications where such properties were
desired. For example, methyl- and ethylmercury compounds were efficiently used globally for
a long time as a cereal seed dressing and also in the paper pulp industry where phenylmercury
acetate was used as a fungicidal additive in the pulp.
Adverse effects on human health associated with organic mercury compounds were
brought to broad public attention in the late 1950s when it was discovered that the widespread
occurrence of neurological disease among residents in Minamata Bay and along the Agano
River, Japan was linked to methylmercury poisoning [5]. Local chemical industries producing
acetaldehyde and vinyl products were eventually pinned down as pollution sources. In the
manufacturing processes used, catalytic inorganic mercury compounds were unintentionally
converted into methylmercury that was discharged with process wastewater to nearby fishing
grounds. The methylmercury accumulated through the aquatic trophic levels and eventually
the people in these areas, who depended heavily on a fish-rich diet, were exposed to
considerable amounts of methylmercury. In 1999 nearly 3000 diagnosed cases of
methylmercury poisoning (Minamata disease) had been recorded in the Minamata-Agano area
[5]. Other severe cases of alkyl mercury poisonings have occurred, for example in Iraq during
the early 1970’s when flour made from methyl- and ethylmercury treated seed-grain was used
to prepare bread. In retrospect, some 40,000 cases of mercury poisonings were observed and
of which approximately 500 were lethal [4].
2.1
Toxicology
The toxicity of mercury is explained by the strong attraction between mercury and sulphur.
When mercury enters a cell it will interact with thiol moieties of amino acids, such as
Cysteine. In this manner protein structure and function, which is largely provided by the
sulphur bridging between thiol functionalised amino acids, is deteriorated. For example, the
enzyme choline acetyl transferase that is involved in the final step of acetylcholine production
is inhibited by mercury. This inhibition may lead to acetylcholine deficiency, contributing to
signs and symptoms of motor dysfunction. The most affected organs for mercury poisonings
are the central nervous system and kidneys. For a detailed summary of biological
abnormalities in connection to mercury exposure see reference 6.
Mercury toxicity is highly dependent on the chemical form, or species, in which it is
present and consequently there are also differences in symptoms developing from exposures
to the three main mercury categories: elemental-, inorganic- and organic mercury. As a result
of the markedly dissimilar chemical and physical properties between mercury categories,
sources of exposure are different. The main exposure source for elemental mercury vapour
nowadays is dental amalgam. For organic mercury compounds, particularly methylmercury,
the major source of exposure stems from consumption of fish and other seafood. This also
applies to inorganic mercury, although some parts of the global population may be subject to
significant exposures through use of mercury in ritualistic/cultural purposes, for example
body decoration with cinnabar based pigments or in traditional medicine.
2.1.1
Elemental mercury
Elemental mercury seldom causes acute toxic effects if orally ingested. The gastro-intestinal
absorption is very low (less than 0.01 %) and the lethal oral dose corresponds to about 100 g
[7]. However, if the vapours are inhaled absorption occurs with near 100 % efficiency. The
2
absorbed elemental mercury is rapidly distributed within the organism and can readily pass
the blood-brain barrier. The elemental mercury is subsequently oxidised to inorganic mercury
via the catalase-peroxide pathway for further interaction with proteins as outlined above.
Chronic mercury vapour exposures affect the central nervous system and kidneys with
symptoms such as tremors, fine motor difficulties and proteinuria and ultimately-death.
2.1.2
Inorganic mercury salts
Owing to the higher water solubility of inorganic mercury salts the lethal oral dose is a mere
0.5 g. However, since inorganic mercury is not lipid soluble it cannot pass the blood-brain
barrier sufficiently enough to cause significant damages to the nervous system. In most cases
of acute inorganic mercury poisonings affected organs include kidney and intestines.
Although rare, mild forms of chronic inorganic mercury are exemplified by the occurrence of
acrodynia or “Pink disease” during the middle of the last century among small children
treated with teething powders containing inorganic mercury salts. Affected persons often
displayed symptoms such as pink extremities and severe leg cramps.
2.1.3
Organic mercury
Organic mercury compounds and in particular methylmercury are nowadays treated as potent
neurotoxins (lethal oral dose is about 0.2 g). Numerous toxicology studies have revealed that
the main target organ for organic mercury compounds is the central nervous system due to
high lipid solubilities that enable efficient penetration of the blood brain barrier as well as the
placental barrier. For methylmercury, absorption occurs at 90-100 % efficiency if orally
ingested. There are marked differences in the toxic actions of methylmercury between the
mature and the developing nervous systems, the latter being most sensitive. In adults, the first
symptom to appear, which can occur after a latency period as long as several month from the
time of exposure, at low dose exposures is usually paresthesia (numbness or a pins and
needles sensation). As the exposure increases, symptoms such as tremors, dysarthria, ataxia,
deafness and constriction of the visual fields occur [8].
The severe toxic actions of methylmercury on the developing nervous system were
first recognized during the Minamata incident outlined above. There it was observed that
women with symptoms of mild methylmercury poisoning gave birth to children with severe
neurological damages that was overall analogous to cerebral palsy. Facilitated by the capacity
to cross the placental barrier methylmercury interferes teratogenically with normal neuronal
development in the foetus and may also have an effect on cell division [9].
In 1999 attention was drawn towards another potential organic mercury poisoning
source when the American Academy of Pediatrics and the US Public Health Service in a joint
statement identified the use of Thimerosal (sodium ethylmercurithiosalicylate) for anti
microbial purposes in medicinal preparations as an extensive source of organic mercury
exposure (ethylmercury) in infants and small children. Eventually, it was hypothesized that
the administration of Thimerosal containing vaccines was associated with neurodevelopmental disorders such as autism and speech delay [6]. In 2001 the US Institute of
Medicine (IOM) concluded that such a relationship was biologically plausible and
recommended that Thimerosal containing vaccines should not be administered to infants,
small children and pregnant women [10].
2.2
The mercury cycle
Around the beginning of the 1960’s, scientists in Sweden were witnessing abnormal
neurological behaviour in predatory birds. As mentioned, alkyl mercury compounds were at
3
the time a widely used agricultural fungicide, and as a result birds on top of the food chain
accumulated high amounts of mercury by preying on small seed-feeding birds and rodents.
Surprisingly, when control groups of fish-feeding birds were analysed, researchers found high
concentrations of mercury in them as well even though these birds had no direct contact with
mercury treated grain seed [11].
Following studies revealed the occurrence of high mercury levels in fish from remote
lakes isolated from point mercury pollution sources. Initially, it was hypothesized that the
ever increasing anthropogenic input of mercury to the atmosphere gave rise to increased
mercury deposition and combined with acid rain caused an elevated mobility and hence
accumulation of mercury in higher organisms within the aquatic compartment [12]. However,
since mercury accumulates through food webs primarily as methylmercury [13] the
atmospherically deposited mercury has to undergo methylation at some stage prior to
bioaccumulation.
In 1969 Jensen and Jernelöv [14] presented evidence that inorganic mercury was
transformed microbiologically to methylmercury in aquatic sediments. This finding combined
with the knowledge gained in Minamata ultimately crystallized into the revelation that people
who were dependent on a seafood based diet could be subjected to methylmercury poisoning
(Minamata disease) just about anywhere. Consequently, in many countries legislative
regulations were imposed to control mercury exposure from ingestion through maximum
allowed mercury concentrations in fish and other seafood. In the European union, the present
limit for mercury in fish is set at 0.5 mg kg-1 for non-predatory species and 1 mg kg-1 in
predatory fish to meet the recommendations issued by the JECFA (Joint FAO/WHO (Food
and Agriculture Organisation of the United Nations and World Health Organisation) Expert
Committee on Food Additives) on a provisional tolerable weekly intake of 1.6 µg
methylmercury per kg body mass [15].
The findings by Westöö and Jensen et al [13, 14] initiated an ongoing effort to unravel
the complex biogeochemical cycling of mercury in the environment as well its toxicology and
involves numerous branches of natural science. The current understanding of the
environmental “mercury-cycle” is displayed graphically in figure 1.
4
Hg0 (g)
Hg2+ (g)
Hg0 (part)
Hg2+ (part)
Hg2+ (aq)
Hg0 (aq)
Atmosphere
CH3Hg+
Hg2+
(CH3)2Hg
Hg0
Water
Hg2+
CH3
Hg0
Sediment
Hg+
(CH3)2Hg
(CH3)2Hg
Hg
Dr
yd
em
ep
iss
os
ion
itio
s
n
s
ion
iss
em
Evasion
Hg
nic
ge
po
ro
th
An
Wet deposition
Hg0
Na
tur
al
Natural Hg runoff
Anthropogenic Hg runoff
Hg2+
HgS
CH3Hg+
Figure 1. Mercury emission sources and transformation mechanisms in the environment.
Elemental mercury is the dominant mercury species in the atmosphere and around
4400 metric tonnes a year is emitted to it from both natural and anthropogenic sources [16].
Natural emissions occur from soil and sea surfaces but also from volcanoes at significant
amounts. Anthropogenic mercury emission sources include combustion of fossil fuels, waste
incineration and mining activities, with the two former examples accounting for 70 % of the
anthropogenic emissions. Being a single atom gas, elemental mercury is highly stable in the
atmosphere with an average residence time of 1 year and can therefore be transported over
very long distances. Mercury is therefore considered a global pollutant in the sense that a
local release can have global effects.
Mercury is deposited to the aquatic and terrestrial compartments with precipitation
following oxidation of elemental- to inorganic mercury. Once the deposited inorganic
mercury enters an aquatic system it may complex with, for example, anions like chloride and
hydroxide or to thiol sites on dissolved or particulate organic material. The various inorganic
mercury species subsequently sediment at bottoms where they either are immobilised by
formation of cinnabar (HgS) or undergo biological transformations into methyl-, dimethyl- or
elemental mercury. Mercury species formed in sediments may thereafter transfer to the water
column where methylmercury is available for further biomagnification in aquatic food webs
while a majority of the dimethyl- and elemental mercury ultimately will emit to the
atmosphere (figure 1).
5
Table 1. Typical methyl- and total mercury concentrations in environmental and biological
matrices, compiled from US EPA data [17]
Environmental media
HgTot
Methylmercury
a
Air
1-170
0-40a
Precipitation
4-90b
0.04-0.6b
b
Fresh water
0.2-15
0.04-0.8b
Sea water
0.3-15b
0.01-0.5b
c
Soil
8-406
0.3-23c
c
Ocean sediments
2-2200
0.06-70c
Lake sediments
10-750c
0.3-30c
d
Fresh water fish
30-330
28-310d
Marine fish
10-1300d
10-1240d
a. ng m-3
b. ng l-1
c. ng g-1 dry weight
d. ng g-1 wet weight
2.2.1
Mercury methylation
In 1985 Compeau and Bartha identified sulphate reducing bacteria (SRB) to be primary
mercury methylators in aquatic sediments [18]. Through a series of inhibition-stimulation
experiments of anoxic estuarine sediment they found that the strain Desulfovibrio
desulfuricans LS displayed very high mercury methylating potential. In following studies
[19], the same research group showed that mercury methylation in Desulfovibrio
desulfuricans LS occurred via donation of a methyl group from methylcobalamin, or
methylated vitamin B12, and further suggested that the reaction was mediated through
enzymatic catalysis in the acetyl-CoA pathway [20]. However, the biochemical pathway for
methylation of inorganic mercury in SRB is still somewhat unclear since methylcobalamine
and acetyl-CoA mechanisms are not at all unique to SRB and it is likely that another, so far
unidentified, component is most likely also involved in the process. Recently it was also
shown that some SRB strains that do not utilize the acetyl-CoA pathway also are capable of
mercury methylation and possibly even without the presence of methylcobalamin [21].
At present there seems to be no absolute answer to why SRB methylate mercury. In
living organisms there are a number of biochemical pathways performing methylation on a
variety of substrates, e.g protective methylation of DNA, and it is likely that mercury
methylation is a coincidental result of inorganic mercury competing with intended substrates
to be methylated. Nevertheless, it has been shown that the methylation of mercury provides
resistance against the toxicity of inorganic mercury in the fungus Neurospora crassa [22].
6
Acetate
Acetyl-CoA
CODH
CO
CH3
CODH
B12-protein
2e-
CH3Hg+
Hg2+
CO2
CH3
Methyltransferase
CH3-THF
4eFormate
FDH
2eCO2
Figure 2. The biochemical pathway for mercury methylation as proposed by Choi et al [20].
Apart from pure culture mercury methylation experiments, the importance of SRB for
the production of methylmercury in different environmental compartments has been further
verified in a number of studies [23-26]. In these, measurements of sediment microbial
community composition and activity in combination with inorganic mercury radiotracers have
shown correlation between the abundance and respiration of SRB populations and rates of
mercury methylation.
SRB have also been found capable of producing dimethylmercury [27]. The rate at
which this process occurs has been shown to be about 1000 times lower than the mono
methylation rate in pure culture as well as in sediments and sediment slurries [28]. Whether or
not the acetyl-CoA pathway is also involved in the mercury dimethylation process is not
clearly established. For instance, Baldi et al [29] suggested that the formation of
dimethylmercury in Desulfovibrio desulfuricans LS originated from the decomposition of
insoluble di methylmercury sulphide ((CH3Hg)2S) formed by this bacteria.
2.2.2
Methylmercury demethylation
A number of bacterial strains, both anaerobic and aerobic, are able to degrade or demethylate
methylmercury (and other alkyl- or phenylmercury species) as a resistance mechanism against
organic mercury compounds. This mechanism is mediated through two different biochemical
7
pathways, reductive- and oxidative demethylation. Reductive demethylation is genetically
controlled via the mercury resistance operon (mer), which is fairly common among microbes
in the environment [30]. Transcription of the mer operon is induced in the presence of
methylmercury resulting in synthesis of organomercurial lyase, an enzyme that cleaves the
mercury-carbon bond in methylmercury yielding inorganic mercury and methane. The
resulting inorganic mercury in turn induces transcription of the enzyme mercuric reductase,
which reduces inorganic mercury to elemental mercury, which subsequently evaporates from
the organism due to the high vapour pressure and low aqueous solubility. Attempts have been
made to incorporate the mer-operon into plants, called transgenic plants, for phytoremediation
of mercury-contaminated soils [31].
NADP+
NADPH
Organomercurial lyase
RHg+
RH
Hg2+
NADPH
Mercuric reductase
Hg0
NADP+
Figure 3. The reductive demethylation pathway as proposed by Robinson and Tuovinen [30].
The oxidative demethylation pathway was recently proposed based on observations of
carbon dioxide instead of the expected methane in the demethylation pathway [32, 33].
Oxidative demethylation is thought to occur within biochemical mechanisms that utilize
single-carbon substrates to derive energy. Presumably, the oxidative demethylation
mechanism yields inorganic mercury as an end product as opposed to elemental mercury in
mer-mediated demethylation. It also appears that the oxidative demethylation pathway is
more dominant in pristine ecosystems whereas the mer-operon is more abundant and hence
reductive demethylation dominates in mercury-contaminated areas [34].
3.
Analytical
speciation
techniques
and
methodology
for
mercury
Mercury speciation analysis can be performed with direct or indirect methods. The direct
methods, involving techniques such as nuclear magnetic resonance (NMR), X-ray absorption
spectroscopy (XAS) and X-ray absorption near-edge structure spectroscopy (XANES), are
highly attractive in the way that samples can be analysed with minimal treatment prior to
measurements. Unfortunately, the direct techniques display relatively poor detection limits,
typically in the low mg kg-1 range, and are therefore not suitable for quantitative mercury
measurements in most samples. Direct techniques are, however, important tools for the
establishment of mercury species chemical ligands and bonding strengths in various matrices
[35, 36].
8
Indirect methods are typically based on hyphenated techniques, a combination
between a separation (chromatographic) technique and a detection technique, for example gas
chromatography-inductively coupled plasma-mass spectrometry (GC-ICP-MS), high
performance liquid chromatography-mass spectrometry (HPLC-MS) and gas
chromatography-atomic fluorescence spectrometry (GC-AFS). For hyphenated techniques it
is required that the sample is pre-treated to permit introduction on the instrumentation. Such
sample pre-treatment often involve a number of operative steps but, on the other hand, it is
possible to achieve very high analyte enrichment in the sample through optimisation of the
treatment procedure. Indirect methods for mercury speciation usually display substantially
lower detection limits (<µg kg-1) compared to direct methods and are hence well suited for
quantitative analysis at trace level analyte concentrations.
3.1
Sample storage, preservation and preconcentration
Mercury species are normally present at very low concentrations in pristine environmental
samples, see table 1, and especially in natural water samples in which, for example,
methylmercury concentrations in the low to mid pg kg-1 range are quite common.
Contamination of samples during sampling is a potential source of error at such low analyte
levels, which requires that clean sampling- and laboratory protocols are employed [37].
In water samples it has been observed that the relative mercury species concentrations
are unstable over extended periods of storage (< 1 month) prior to analysis as a result of
biological transformation and/or adsorption to sample container surfaces [38, 39]. For short
term storage (< 1 week) these effects can be subdued by storing samples in containers made
of inert materials like polytetrafluoroethylene (PTFE) at + 4°C but obviously it is desirable to
maintain the original species distribution more or less permanently. Some studies have been
concerned with water sample preservation by means of acidification with mineral acids [40,
41] or complex binding [42] of methylmercury on additives such as dithiocarbamate
functionalised resins followed by elution under acidic conditions. However, the subsequent
treatment steps usually require less acidic conditions that must be restored with additional
reagents, which potentially can increase blank levels. Freeze preservation of water samples at
–80 ° C has been investigated with positive results [39, 43]. For biological and sediment
samples the problem with analyte instability is less pronounced. Increased analyte stability in
these matrices might be explained by the high content of reduced sulphur groups on which
mercury species are immobilised. Storage in a regular household freezer dedicated to trace
level samples at –20 ° C is normally a sufficient preservation strategy for these types of
samples.
Sub ng kg-1 levels of mercury species found in natural waters necessitate extraction
and preconcentration of the analytes regardless of detection technique used. To accomplish
this various preconcentration techniques such as distillation [44, 45], solid phase extraction
(SPE) [46, 47] and solid phase micro extraction (SPME) [48, 49] can be employed. In the
author’s opinion though, purge-trap techniques based on aqueous phase derivatisation
combined with adsorbent trapping are best suited for this purpose given that analyte
derivatisation and extraction is carried out in a single step with very high preconcentration
capacity [39, 50-52, I, II, V]. The required equipment is relatively simple and adaptable for
field use if so desired. The technique also offers great sample volume flexibility, which can be
an important feature in order to obtain satisfactory method detection limits when dealing with
trace level analyte concentrations. Using the purge-trap methods for mercury species
derivatisation and preconcentration in aqueous samples is not problem-free though. It has
been observed [39, 52] that aqueous phase ethylation, using sodium tetraethylborate
(NaB(C2H5)4), in aqueous samples containing high concentrations of halides and/or DOC can
9
result in reduced analyte recoveries or transformations of methylmercury, deteriorating
detection limits and increasing the risk for systematic errors. As a consequence,
methylmercury determinations in sea- and mire waters with purge-trap methods based on
aqueous phase ethylation usually require analyte-matrix separation by repeated liquid
extractions under acidic conditions prior to derivatisation to boost analyte recoveries.
Unfortunately, such extraction procedures require the use of several solvents and reagents,
which will increase the blank levels and deteriorate detection limits.
For solid samples such as soils, sediments and biota a majority of current sample
treatment strategies for mercury species extraction are in general variations on the original
Westöö-method, developed by Swedish scientist Gunnel Westöö in 1966 [53]. These
solid/liquid (or liquid/liquid) extraction methods involve either a complete dissolution or
leaching of the samples under acidic or alkaline conditions to liberate mercury species
attached to reduced sulphur groups present in the matrix. The analyte is thereafter separated
from the matrix by liquid extraction using a suitable organic solvent. A variety of leaching
media are used, for example diluted hydrochloric acid, mixtures of sulphuric acid/potassium
bromide/copper sulphate or methanolic potassium hydroxide (KOH) solutions. Sample
digestion methods used for total mercury concentration measurements, for example
pressurised microwave digestion in concentrated mineral acids, are in most cases not well
suited for speciation analysis since such treatment can cause substantial degradation of
organic mercury species. Complete sample dissolution under the relatively mild treatment
conditions required to maintain the species distribution integrity is typically only achievable
with biological matrices. For this purpose, samples can be treated with aqueous solutions of
tetramethyl ammonium hydroxide (TMAH) or KOH (20-25 % w/v) assisted with mechanical
or ultrasonic shaking.
In addition to traditional solid-liquid or liquid-liquid extraction methods, more “up to
date” sample preparation techniques like accelerated solvent extraction (ASE) [54],
supercritical fluid extraction (SFE) [55, 56] and open focused microwave digestion [57] have
been tested in a variety of environmental and biological matrices. However, the techniques
mentioned have not enjoyed widespread acceptance either because analyte recoveries
obtainable are insufficient or that instrument production has been cancelled.
3.2
Mercury species separation
To perform a speciation analysis using an indirect method the mercury species present in the
processed sample need to be separated in time prior to detection. Initially in the 1960s, gas
chromatography (GC) was the preferred separation technique in mercury speciation science.
Today virtually all known chromatography techniques have been evaluated and applied for
mercury speciation. All contemporary chromatographic techniques have their pros and cons
but, in the author’s opinion, gas chromatography still stands out as the most attractive to apply
in the field of mercury speciation on account of unparalleled separation efficiency and
robustness.
3.2.1
Gas Chromatography
In gas chromatography (GC) the sample is normally introduced as an organic solution in an
injector, where it is vaporized and transferred by the mobile phase (gas) through a
chromatographic column situated inside a heated oven. The GC-oven can be either held at a
constant temperature (isothermal) or programmed to increase the temperature at a given rate.
Analytes are separated on basis of differences in volatility and, for some applications, polarity
between analyte compounds and their interaction with the column stationary phase. Today,
most GC separations are carried out on capillary columns, with inner diameters of 0.25 to
10
0.53 mm, 15-60 m of length and stationary phase thickness between 0.25-3 µm. The
stationary phase consists of a polysiloxane polymer that is characterized by a repeated
siloxane backbone in which each silicon atom contains two functional groups. The amount
and type of functional groups are varied to obtain different stationary phase properties, for
instance 100 % dimethyl polysiloxane is a non-polar phase whereas the addition of 50 %
cyanopropylmethyl polysiloxane or trifluoropropylmethyl polysiloxane gives a high polarity
phase.
CN
CH2
CH3
O
CF3
CH2
CH2
Si
CH2
CH2
O
CH3
O
n
Si
Si
CH3
CH3
n
n
Figure 4. Polysiloxane polymers in capillary GC column stationary phases: Dimethyl
polysiloxane, cyanopropylmethyl polysiloxane and trifluoropropylmethyl polysiloxane
For mercury speciation capillary column GC with non-polar or slightly polar
stationary phases (e.g. 5 % diphenyl polysiloxane phases like DB-5 and other equivalent
phases) offers fast analyses (3-5 minutes) with excellent chromatographic performance that
helps obtaining low detection limits because chromatographic peaks can be highly
compressed in narrow bandwidths, thereby resulting in improved signal to noise ratios.
However, a disadvantage with GC compared to other chromatographic techniques, e.g.
HPLC, is that ionic analytes, such as methylmercury and inorganic mercury, are difficult to
separate regardless of stationary phase used (non-polar or polar). To circumvent this, ionic
species are typically transformed into volatile and stable derivatives through nucleophilic
alkyl substitution by adding a derivatisation reagent to the sample. Halogenated alkyl
magnesium compounds, called Grignard reagents (after the inventor Victor Grignard), and
tetra-alkylated borates like tetraethylborate or tetrapropylborate are the most popular
derivatisation reagents in mercury speciation. Grignard reagents are very unstable in aqueous
media and therefore the analytes have to be derivatised in a dry (water free) organic solvent,
such as toluene or hexane. It is therefore necessary to extract mercury species present in
aqueous sample solutions into such a solvent. This can be accomplished by adding a
complexing agent to the sample, for example diethyldithiocarbamate (DDTC). Tetra-alkylated
borate derivatisation reagents on the other hand are relatively stable in both aqueous and
organic solutions, which means that these derivatisation reagents can be added directly to
aqueous sample solutions. The alkylated non-polar mercury species can thereafter either be
11
extracted into an organic solvent, for traditional syringe injection on the GC, or be purgedtrapped on adsorbent columns as discussed in section 3.1.
3.3
Mercury species detection
Traditionally, electron capture detectors (ECD) have been used extensively in combination
with GC separations of mercury species. This type of detector is selective towards
electronegative compounds, especially halogenated molecules, and therefore mercury species
have to be derivatised towards such functionality when the ECD is used. Nowadays ECD
detectors have been more or less abandoned in the field of mercury speciation mainly due to
the risk that other halogenated (electronegative) compounds co-elute with the mercury species
under study, thereby producing erroneous results. Instead spectrometric detectors that display
high selectivity towards mercury are preferred, e.g. atomic absorbtion spectrometry (AAS),
atomic fluorescence spectrometry (AFS), atomic emission spectrometry (AES) and mass
spectrometry (MS), since these detection techniques greatly reduce the number of possible
matrix interferences. The work presented in this thesis was carried out using inductively
coupled plasma quadropole mass spectrometry (ICP-QMS) as a GC detector (GC-ICP-QMS).
Several benefits can be gained by using mass spectrometric detection for mercury speciation
work. The ability to discriminate between elemental isotopes is particularly useful as it
permits unique experiments to be carried out.
3.3.1 ICP-MS
In mass spectrometric techniques, molecular or elemental ions produced in an ion source are
sorted and counted in a mass analyser on basis of their mass to charge ratio (m/z). There are
several different mass analyser designs, for example quadropole-, time-of-flight-, ion trapand sector field mass analysers that display varying capacities to resolve different mass to
charge ratios. For characterisation of inorganic/organometallic compounds mass spectrometry
can be used with common ion sources, such as electron ionisation (EI) and electro spray
ionisation (ESI). However, for quantitative elemental analysis of such compounds a complete
decomposition of the sample is required to avoid severe spectral interferences (isobaric
overlap). To accomplish this, “hard” ion sources that atomise and ionise the analytes with
high efficiency, like thermal ionisation, glow discharge and inductively coupled plasma, are
applied.
Among hard ion sources, the inductively coupled plasma is the most suitable to use for
elemental speciation analysis since it is compatible with continuous sample introduction and
therefore more easily coupled to chromatographic systems like GC and HPLC. It consists of
three concentric quartz tubes, called a torch, through which streams of Argon are passed at a
combined flow rate of typically 15 l min-1. The top of the torch is surrounded by an induction
coil that is connected to a radio frequency generator operating at either 27 or 40 MHz with
about 1500 W power capacity. The argon gas is initially ionised by introducing Tesla sparks
into it and an Argon-plasma is formed at the top of the torch. The plasma is sustained by the
collisions between Argon atoms and electrons in the oscillating inductive field. At the high
frequencies used plasma temperatures up to 10000 ° C are obtained locally, which is
sufficient to produce singly charged atomic ions at high efficiency.
12
Ions produced in the ICP ion source are extracted into the mass spectrometer through a
source/spectrometer-interface that sequentially lowers the pressure from atmospheric to about
10-4 Pa by differential pumping. After passing the interface analyte ions are then focused and
accelerated by a series of ion lenses into the quadrupole mass analyser or mass filter, which
separates the ions based on their mass/charge ratio. The quadrupole type of mass analyser is
built up of four parallel hyperbolic stainless steel rods to which a combination of RF and DC
voltages are applied. For a specific combination of these voltages only ions of a specific
mass/charge ratio will have a stable flight path through the mass analyser to the detector,
enabling selective mass filtering. Filtered ions are subsequently counted in the detector
(electron multiplier). The high ionisation efficiency of the ICP source and combined with a
clean background spectrum generate very high signal to noise ratios and hence low detection
limits for most elements.
Interface
Plasma gas
Quadropole
Plasma
Auxilary gas
Sample
M+
Detector
Induction coil
Torch
Ion lenses
Ion source
Mass analyser
Figure 5. Schematic presentation of ICP-MS (M+ signifies analyte ion)
3.3.2
GC-ICP-MS
Couplings between capillary column GC and ICP-MS were first described almost 15 years
ago [58]. The GC-ICP-MS technique has quickly increased in popularity and is nowadays a
frequently used technique for speciation of most metals and metalloids but also for halides or
halogenated hydrocarbons as well as organic phosphor and sulphur compounds. The main
benefit of using GC-ICP-MS for mercury speciation is a significant gain in sensitivity
compared to other mass spectrometric techniques such as GC-MS. In general, detection limits
1-2 orders of magnitude lower than those of GC-MS can be obtained, which of course is
highly valuable when dealing with trace analyte concentrations, e.g. natural water samples. A
major drawback of the technique, in addition to high running costs (Argon consumption at
~15 l min-1), lies in the fact that no molecular weight or structure information for polyatomic
analytes can be extracted as a result of the complete analyte decomposition that is effected in
the ICP ion source. Identification of individual mercury species therefore has to rely upon
retention time matching against high purity standard compounds.
13
3.4
Isotope dilution mass spectrometry (IDMS)
The isotope selectivity of a mass spectrometric detector makes it possible to employ so called
isotope dilution calibration to mercury speciation analysis. The general principle of isotope
dilution methodology is that the sample is standardised by adding a known amount of a
compound (commonly called “spiking”) that has identical chemical properties as the analyte
apart from being labelled with an enriched isotope (isotopomer). This results in an
experimental isotopic abundance ratio between isotopomer and incipient analyte (reference)
that deviates from the naturally occurring ratio (figure 6).
Intensity
Sample
m/z
B
A
Intensity
Isotopomer standard
A
B
m/z
B
m/z
Intensity
Mixture
A
Figure 6. Isotope dilution demonstrated by an analyte having two isotopes with m/z A and B
in the sample, isotopomer standard and spiked sample
Based on the measured experimental isotope ratio in the mixture between reference
and isotopomer in the sample as well as corresponding ratios in the unspiked sample and
isotopomer standard, the analyte amount can be calculated according to equation 1.
C
x
 C sm sM
= 
 m xM x
s
 A S − R m B
 
 R m B x − A
14
s
x



(1)
In this equation Cx and Cs denote the analyte concentration in the sample and the
concentration of the isotopomer standard and ms and mx are the masses of the added spike and
sample, respectively. Ms and Mx are the relative molar masses of the spike and analyte. Rm is
the measured ratio between reference (A) and isotopomer (B) in the spiked sample. As and Bs
are the atom fractions of reference and isotopomer in the isotopomer standard, while Ax and
Bx are the equivalent for the original sample. As can be seen, the analyte concentration in the
sample can be determined without the establishment of calibration curves, in one single
measurement of Rm. All other parameters are given in the isotopomer standard specifications
and from weighing the sample and spike. However, in the case of the ICP-MS technique used
in this thesis, analyte ions with high m/z are detected at higher relative sensitivity compared to
lower m/z analyte ions, which means that measured ratios are distorted. This so called mass
bias effect is compensated for in the calculation of results by applying experimentally
determined correction factors or abundance values, which requires additional analyses.
Since isotope dilution quantitation is based on isotope ratios that remain constant
during the full analysis cycle (sample pre-treatment, pre-concentration, analysis), potential
error sources such as analyte degradation and detector drift etc will not influence the final
result since analyte and standard are affected to an exact equal extent. Provided that
isotopomer and analyte extraction yields are equal, the accuracy and precision that can be
obtained by using IDMS at optimum conditions is typically superior to external or regular
internal calibration approaches. Equally important is the fact that, when using IDMS, nonquantitative analyte recoveries will not have an effect on the accuracy of results (within
reasonable limits), which ultimately means that sample pre-treatment procedures can be
significantly simplified and overall sample throughput improved.
IDMS has been an established quantitative method for quite some time in organic
mass spectrometry, for example for the determination of PCBs and other persistent organic
pollutants (POPs) in a wide range of matrices by using 2H or 13C labelled isotopomers. The
application of IDMS in speciation analysis (usually referred to as “species-specific isotope
dilution” (SSID)) was pioneered by Heumann some 15 years ago parallel to the initiation of
the GC-ICP-MS technique [59, 60]. To date SSID based speciation methods for the
determination of selenium, chromium, thallium, lead, iodine, tin and mercury have been
published. Contrary to the isotopomers used in organic IDMS, the isotopic labelling of the
tracer compounds used in speciation IDMS is usually incorporated in the parent trace element,
for example CH3200Hg+ rather than C2H3Hg+ or 13CH3Hg+, to maintain high sensitivity at
higher atomic mass/charge ratios with ICP-MS detection and also to allow for SSID
calibration of valency species, inorganic mercury (Hg2+) for example. However, since this
approach for isotopic labelling of trace element species requires that the parent trace element
holds at least two stable isotopes (polyisotopic), SSID speciation methods for monoisotopic
elements such as Arsenic would have to be based on long lived radioisotopmers or
deuterated/13C labelled isotopomers where applicable.
Table 2. IUPAC mercury isotopic abundances (atom %).
Isotope
196
Abundance
0.15
Hg
198
Hg
9.97
199
Hg
16.87
15
200
Hg
23.10
201
Hg
13.18
202
Hg
29.86
204
Hg
6.87
No method is perfect and of course this also applies to isotope dilution calibration.
The precision of ID-based measurements is dependent on the ratio between reference and
spike isotopes, A and B in figure 6. For any given system of spike/reference isotopes there is
an optimum ratio between A and B at which the optimal precision in measurements is
obtained [61]. Ratios higher or lower than the optimum increases the uncertainty of the
measurement to a proportion that is determined by a so-called error propagation factor (EPF),
which in turn is derived from the isotopic abundances of the spike and reference isotopes in
both the enriched standard and natural sample [62]. To achieve the optimal ratio when adding
the isotopomer standard it is therefore necessary to know the concentration of the analyte
present in the sample. In reality, however, or at least in the case of mercury, the detrimental
effect caused by non-optimal isotope ratios on the overall measurement uncertainty is
relatively small compared to contributions from other uncertainty sources [V]. An estimation
of the incipient analyte concentration based on previous results or results presented in the
literature for similar matrices can in many cases be adequate to obtain reasonably optimal
ratios between spike and reference isotopes. Figure 7 describes the EPF as a function of 200Hg
to 202Hg isotopomer/reference isotope ratios and as can be seen the ratio can be varied within
a quite wide range without causing dramatic increase in the EPF.
1.7
Error propagation factor
1.6
1.5
1.4
1.3
1.2
1.1
1
0
5
10
15
20
25
30
35
40
45
200Hg/202Hg
Figure 7. Error propagation factor as a function of 200Hg/202Hg ratio.
3.4.1
Isotopomer synthesis
SSID calibration is a recently established methodology in speciation analysis and the number
of laboratories that apply it are relatively few, therefore isotopomer standards or certified
reference materials (CRMs) have, up until recently [IV], not been commercially available. As
a result scientists in the field have had to synthesise their own isotope enriched standards. In
the case of mercury speciation, aqueous isotope standards of inorganic mercury can easily be
produced by dissolving the enriched mercury isotope (usually supplied as oxide or as the
metal) in nitric acid followed by dilution with doubly distilled (or better) water to desired
concentrations. For the preparation of methylmercury isotopomer standards three approaches
for synthesis routes have developed. The most popular seems to be to use methylcobalamine
[52, 63] or tetramethyltin ((CH3)4Sn) [64-66] to accomplish a mono methyl group transfer to
16
isotope enriched inorganic mercury in aqueous media with high synthesis yields. These two
similar synthesis routes have the advantage that methylmercury can be synthesised without
significant amounts of residual mercury species forming, like dimethylmercury or inorganic
mercury. Consequently, subsequent product purification procedures are drastically simplified
or in best case unnecessary. For the work presented in this thesis a different synthesis strategy
based on a so called comproportionation reaction [67], was applied in the production of
mercury species isotopomer standards. This synthesis route initially produces
dimethylmercury quantitatively by reacting mercuric chloride with a Grignard reagent,
CH3MgCl, in organic solution (toluene). Mono methylmercury chloride is subsequently
produced by refluxing equimolar amounts of dimethylmercury and mercuric chloride. This so
called comproportionation reaction yields mono methylmercury chloride at near 100 %
efficiency. However, the handling of dimethylmercury can be extremely hazardous if dealing
with concentrated solutions and appropriate safety precautions must always be taken [68].
The largest benefits of this method for isotopomer synthesis are that high yields are obtained
and that it is equally applicable to synthesis of other alkylated or phenylated mercury species,
such as ethylmercury chloride [III], simply by adjusting the type of Grignard reagent.
In cases when a speciation analysis concerns more than one element species, SSID
calibration can be applied as a quantitative method in either “single labelled”- [67, 69, 70] or
“multi labelled” mode [III, 71, 72]. In single labelled SSID isotopomers for all element
species under study are labelled with the same isotope and correspondingly, in multi labelled
mode, each species-specific isotopomer is labelled with an individual isotope. Multi-labelled
SSID speciation analysis is highly useful since it facilitates detection and subsequent
correction [III, 73] for possible species interconversion reactions, or transalkylation between
species and matrix occurring during sample storage and pre-treatment that in some cases can
introduce severe bias in the quantitative analysis. This can be exemplified by the unintentional
formation of methylmercury during sample processing of sediment samples that occur when
steam distillation techniques are used for analyte extraction/pre-concentration. The steam
distillation technique became a subject of debate in the late 1990s when Hintelmann and coworkers found by spiking sediment samples with an inorganic mercury isotopomer followed
by distillation enrichment and GC-ICP-MS analysis, that methylmercury was produced during
the distillation processing [74]. The degree of such “artefact formation” of methylmercury
was found to be around 0.03 % of the added inorganic mercury spike. The ratio between
inorganic to methylmercury concentrations in marine sediments normally is about 100:1 [75]
and therefore a small unintentional formation of methylmercury during sample processing
would not influence the analysis result significantly. On the other hand, in contaminated
sediments this ratio can sometimes be as high as 1500:1 and in such samples even a 0.03 %
“artefact formation” of methylmercury would result in a substantial over estimation of the
methylmercury content. Studies aimed at understanding the mechanisms leading to so-called
“artifact formation” of methylmercury during distillation sample processing have showed that
it is related to the inorganic mercury content of the sample and that it can be kept very low or
insignificant if conventional solid/liquid extraction methods are employed instead of
distillation in the processing of sediment samples [I, II, 74, 76-79].
3.5
Analytical methods for monitoring mercury species transformations
in the environment
It is nowadays established that observed methylmercury concentrations in aquatic sediments
is the net product between mercury methylating and methylmercury demethylating
mechanisms effected by micro organisms. Consequently, it is of interest to gain knowledge
about the rates at which these reactions occur as well as their environmental controls. Such
17
information can be valuable in, for example, risk assessment of mercury polluted industrial
areas and the development of preventive measures to minimise mercury methylation and
emission at such sites.
The methods by which mercury species transformation rates have been determined in
sediments involve three different strategies. Jensen and Jernelöv [14] first studied mercury
methylation in aquatic environments by spiking sediments with inorganic mercury of natural
isotopic composition. In most cases when using this method the sediments have to be
amended with very high concentrations of inorganic mercury, several µg g-1, to differentiate
any significant methylation of the added inorganic mercury spike from the methylmercury
already present in the sample. Unless the sediment under study originates from a heavily Hgpolluted site, the results obtained could be considered uncertain since large inorganic mercury
additions will probably affect the incipient mercury species equilibria and microbial fauna
drastically. More specific methods that utilize radiochemically labelled tracers, e.g. 203Hg2+
[80] and 14CH3Hg+ [81], have been used frequently for mercury methylation as well as
methylmercury demethylation studies. In the radiotracer methods the amount of methylated
radioactive 203Hg2+ in the sediment is typically quantified by selective methylmercury
extraction followed by scintillation counting. The amount of demethylated 14CH3Hg+ can be
estimated by collecting the headspace of the incubation vessel and measuring the radioactive
methane (14CH4) or carbon dioxide (14CO2) formed in the reductive or oxidative
demethylation pathways. Radiotracer methods are appealing because the required
instrumentation is relatively simple and methods can easily be adopted for field use.
However, the radioactive 203Hg2+ tracer is usually of low specific activity and therefore
samples have had to be spiked to unrealistically high concentrations to detect the methylated
tracer. Gilmour and Riedel [82] used a custom-made 203Hg2+ tracer with very high specific
activity to circumvent this dilemma, which allowed them to determine mercury methylation
rates in sediments with very small tracer additions (<10 % of incipient levels). Nevertheless,
the selectivity of the methylmercury extractions used in radiotracer methods is questionable,
meaning that there is a risk for residual 203Hg2+ to be co-extracted together with the
CH3203Hg+ and thereby methylation rates may be overestimated. Demethylation studies using
the radiotracer methodology are also problematic since a given proportion of the end products
in the demethylation pathway, 14CH4 and/or 14CO2, are most likely reincorporated into the
microbiological biomass, leading to underestimated demethylation rates.
Hintelmann and co-workers significantly improved the methodology for mercury
species transformation studies about ten years ago when they presented a method that made
use of enriched stable inorganic mercury as a tracer to monitor mercury methylation in
combination with GC-ICP-MS instrumentation [83]. By spiking sediment slurries with
199
Hg2+ at sub-incipient concentrations, mercury methylation rates as well as changes in
incipient methylmercury levels could be determined over a period of three weeks. They
further developed their method to also include a methylmercury tracer, which enabled
measurement of demethylation rates simultaneously in the same sample as methylation rates
were measured [84, 85].
The general theory of using enriched stable mercury tracers for this type of studies
resembles isotope dilution calibration methodology. If an isotope enriched inorganic mercury
tracer is added to a sediment, for example 199Hg2+, the isotopic abundance ratio between 199Hg
and a reference isotope (normally 202Hg) for that species will be shifted from the natural ratio
to a higher value that is dependent on the amount of added tracer. Methylmercury produced
from the inorganic mercury present in the sample will then also display a 199 to 202 ratio that
is higher than the natural 199/202 ratio. The amount of produced CH3199Hg+ is proportional to
the degree that the measured methylmercury 199/202 ratio exceeds the natural 199/202 ratio
and can usually be quantified with high accuracy and precision, especially if the tracer
18
experiment is combined with SSID calibration during sample treatment. The limit of detection
for mercury methylation using the stable isotope methodology will be dependent on the
precision of the isotope ratio measurements and the incipient concentration of methylmercury.
To detect a significant methylation of the added 199Hg2+ tracer, the resulting increase in the
199/202-methylmercury ratio has to exceed the natural 199/202 ratio by an extent equivalent
to three times the standard deviation interval at which the 199/202 ratio can be measured. In
line with this, the incipient methylmercury concentration also determines the methylation
limit of detection in each individual sample, as a higher incipient amount of methylmercury
means that a higher amount of the tracer has to be methylated to produce a significant
increase in the 199/202 methylmercury ratio and vice versa. In that respect, radiotracer
methods are advantageous since the added 203Hg2+ tracer does not exist naturally and hence
does not suffer from mercury related spectral interferences. The precision at which isotope
ratios can be measured with a ICP quadropole mass spectrometer in continuous signal mode is
normally in the neighbourhood of 0.1-0.2 % relative standard deviation (RSD) [86] but
deteriorates if chromatography is involved, like in the case of GC-ICP-MS instrumentation,
since measurements then are based on transient signals. Precision in transient signal
measurements can be improved by using, for example, multi collector ICP-MS (MC-ICP-MS)
instrumentation that is capable of measuring isotope ratios at parts per million level RSD [87].
The strategy for methylmercury demethylation rate measurements using stable
isotopes is different compared to radiotracer methods. Instead of targeting the end products of
the demethylation pathway demethylation rates are indirectly measured based on the decrease
in concentration of the added methylmercury tracer during the incubation period. The
reductive and oxidative demethylation mechanisms produce different end products, Hg0 and
Hg2+ respectively, of which the former is lost to the head space/atmosphere and hence escapes
detection. Furthermore, the fact that methylmercury is generally present at 1-5 % of inorganic
mercury concentrations would render very high demethylation detection limits if
measurements were based on an increase of the inorganic tracer/reference isotope ratio due to
methylmercury tracer demethylation. For an in-depth description on the calculation scheme of
methylation and demethylation detection limits using the stable isotope methodology, see
reference 84. In summary, the stable isotope approach is a very convenient approach for
measurements of mercury species transformations in sediments and other ecological or
biological systems (see figure 8 for a schematic presentation of both radiotracer and enriched
stable isotope tracer methodology for measurement of mercury species transformation rates in
sediment). The greatest benefits gained in using the stable isotope tracer methodology is that
methylation- and demethylation rates can be determined simultaneously in one individual
sample with high specificity and with very low tracer additions, which enables the
establishment of “true” net methylation rates for the system under study.
19
Radioisotope tracers
N2
CH3198Hg+(aq)
Sediment core
Sediment core
203Hg2+(aq)
Stable isotope tracers
14CH Hg+(aq)
3
Anaerobic incubation
N2
201Hg2+(aq)
Anaerobic incubation
CH3200Hg+(aq)
199Hg2+(aq)
CH3203Hg-extraction
Scint.
Counter Headspace
Extraction
GC-ICP-MS
14CO , 14CH
2
4
+ High sensitivity
+ Incipient MeHg concentration,
methylation-/de-methylation rates and
“artifact MeHg” in one experiment.
+ IDMS
- Complex instrumentation
- Low sensitivity
- Selective MeHg extraction?
- CO2 and CH4 can be incorporated into
the microbial biomass
- No incipient MeHg data extractable
+ Simple instrumentation
Figure 8. Comparison between radioisotope- and stable isotope methodologies for monitoring
mercury methylation and methylmercury demethylation in sediments.
4.
Results: Summary of published papers and manuscripts
4.1
Mercury methylation and methylmercury demethylation in brackish
water sediments: Improved methodology and geochemical controls,
papers I and II
In paper I a method for simultaneous monitoring of mercury methylation and methylmercury
demethylation rates in brackish water sediment based on enriched mercury isotope tracers was
developed. In addition, by adding enriched isotope standards of methyl- and inorganic
mercury to the sediment prior to sample treatment, it was also possible to determine both the
incipient methylmercury concentrations by SSID calibration with a precision of 2-3 % RSD
as well as the amount of artefact methylmercury produced in the sample treatment.
Methylmercury was acid leached from the sediments using a mixture of KBr/CuSO4/H2SO4
and extracted to an organic phase (CH2Cl2) followed by back-extraction of the organic phase
to water and subsequent aqueous phase ethylation using sodium tetraethylborate. It was found
that this sample treatment resulted in an average methylmercury “artefact formation” of about
0.05 %. It was also found, in line with observations made by Demuth et al [52], that
methylmercury was partly degraded and reduced to elemental mercury during the sample
20
treatment cycle and/or analysis as the specific isotopic pattern displayed for methylmercury
was also found in the peak corresponding to elemental mercury. Another peak that displayed
a similar isotopic pattern eluted shortly after diethylmercury (inorganic mercury) but the
compound could not be identified. However, when the solid/liquid extraction was performed
in diluted HCl instead of the KBr/CuSO4/H2SO4 mixture this peak was not present and
therefore it was suggested that the unidentified peak might have resulted from methylmercury
bromide. Although the developed method obviously was prone to species transformations
during sample treatment and analysis accurate results were still obtained as a result of the
SSID calibration.
The developed method was applied to assess methylation and demethylation rate
profiles in brackish water sediment from the Öre estuary located in the Gulf of Bothnia.
Measurements were carried out by spiking 1 cm slices of the top 10 cm of a sediment core
with methyl- and inorganic isotope enriched tracers (CH3198Hg+ and 201Hg2+, respectively) to
about 20 % of the incipient methyl- and inorganic mercury concentrations followed by
anaerobic incubation in a nitrogen atmosphere for 10 days. From the obtained results a
corresponding net methylation profile was derived (figure 9). The highest observed net
methylation occurred in the 2-3 cm depth region of the sediment core in which a net
methylmercury amount of about 3.5 ng per gram of dry sediment was produced during a 10day anaerobic incubation. This observation was in line with other studies that also have
described high methylation rates in the top section of lake- and sea sediments [88, 89].
Logically, the net methylation depth profile in sediments ought to reflect the vertical depth
distribution of mercury methylating micro organisms (SRB). For instance, Devereux et al [90]
were able to correlate high near-surface mercury methylation rates to SRB population density
in estuarine sediment by using oligonucleotide molecular probes to determine SRB
community depth distribution.
Incipient methylmercury concentrations varied as much as 5 times over the sampled
10 cm depth and given that the methylmercury levels could be determined with high
precision, even small differences in concentration could be detected (figure 9). Similar to the
net methylation profile, incipient methylmercury concentrations peaked in the 2-3 cm
segment at around 5 ng g-1 but also displayed a second increase at 6-8 cm depth below the
sediment surface with concentrations of about 5.5 ng g-1. The second methylmercury
maximum coincided with a sharp increase in sediment pore water sulphide concentration,
which was measured in the sediment core using a sulphide specific microelectrode, and
therefore the increase in methylmercury concentration at the 6-8 cm depth was possibly due to
immobilisation of methylmercury on reduced sulphur.
21
Net methylation (ng/g dry weight)
Total Hg (ng/g); Total Sulfide (µM); O2 Saturation (%)
0
20
40
60
80
100
120
0
0
1
1.5
2
2.5
3
3.5
4
0
1
1
2
2
3
3
Depth (cm)
Depth (cm)
0.5
4
5
6
7
Total Hg
4
5
6
7
Total Sulfide
8
8
O2 saturation
9
Ambient MeHg
9
10
0
1
2
3
4
5
10
6
Ambient methyl mercury (ng/g)
Figure 9. Depth profiles of sulphide-, oxygen-, total mercury and methylmercury
concentrations in addition to corresponding net mercury methylation rates observed in a
brackish water sediment
A number of factors have been identified to influence microbial mercury methylation
processes in aquatic sediments, e.g. temperature [88], salinity [91], sediment sulphate- and
sulphide concentrations [92-94] and sediment organic material content [88, 95, 96]. In
contrast, basically nothing is known about how by these factors affect methylmercury
demethylation processes.
In paper II the effects of organic material (OM) content and sulphide concentration on
methylation and demethylation processes as well as incipient methylmercury concentrations
were investigated in Öre River estuary sediments by using the method developed in paper I.
Samples were collected from five different sites (erosion- or accumulation type bottoms) in
October, May and June 2001-2002 to incorporate a broad spatial and temporal variation in
sediment OM- and sulphide content into the experimental design (table 1 and figure 1 in
paper II).
The measured incipient methylmercury concentrations and mercury methylation rates
varied considerably with bottom type, sediment depth and sampling occasion (figure 2 in
paper II). In general, sediments that contained higher levels of OM also displayed relatively
high methylmercury concentrations compared to low OM containing sediments. In the
accumulation bottoms, methylmercury concentrations comparable to those normally found in
Hg-contaminated marine environments were observed, which was remarkable since the Öre
estuary is considered a pristine area [97]. This finding indicated that process-controlling
factors, other than the availability of substrate (Hg2+), are important for microbial synthesis of
methylmercury in sediments unpolluted with mercury. The correlation between sediment
organic material and methylmercury concentrations was found to be significantly correlated
in an exponential regression model based on the total set of data (figure 3 in paper II). With
respect to temporal variation in incipient methylmercury concentrations, the highest levels
were observed in October and the lowest in May. Mercury methylation followed the same
spatial pattern and varied by factor of ~10 between sites. Temporarily, the highest mercury
methylation rates were recorded in June or October and the lowest in May in all sites. In most
cases the depth of maximum methylation rates for individual sampling sites were similar
between sampling occasions. Demethylation depth profiles were, however, surprisingly
22
constant spatially as well as temporarily. Consequently, the calculated net methylation depth
profiles were to large degree determined by the mercury methylation potential at all sites
except for a shallow, low OM, erosion bottom that almost exclusively displayed negative net
methylation rates.
To further evaluate how the variables: OM-content, sulphide concentration, mercury
methylation and methylmercury demethylation correlated with incipient methylmercury depth
profiles, partial least squares regression modelling (PLS) was applied to each site and
sampling occasion. In a majority of cases the PLS models explained >85 % of the variance in
incipient methylmercury concentration although the distribution of significant regression
coefficients between models was mixed (table 4 in paper II). However, a PLS model based on
the total data set explained 69 % of the variance in incipient methylmercury with OM-content
and mercury methylation as positive significantly correlated regression coefficients.
Combined, the results indicated that sulphide concentrations in the sediments were in
most cases to low to inhibit mercury methylation, which has been reported by others to occur
at >10 µM sulfide concentrations [92]. Furthermore, methylmercury demethylation processes
in the sediments were not markedly affected by changes in OM- or sulphide concentrations.
The observed spatial variation in incipient methylmercury concentrations as well as depth and
rates of maximal methylation were controlled by the organic material content and differences
in sediment redox conditions. Based on the results obtained in paper II and data reported by
others (table 5 in paper II), it was concluded that non Hg-polluted brackish environments
might constitute potential methylmercury “hotspots”, which may be activated if organic
material loads increase.
4.2
Mercury species transformations in Thimerosal treated mice
monitored with multi labelled SSID, paper III
Suspicion that Thimerosal (ethylmercurithiosalicylate) as a preservative in vaccines used in
child immunisation programmes is linked to increasing numbers of children diagnosed with
neuro developmental disorders (particularly autistic spectrum disorder, ASD) was raised by
Bernard et al [6] in 2001. Since then a number of epidemiological research studies have either
confirmed [98] the suspicion or have not found any correlation at all [99]. Even though
conclusive evidence that confirms a Thimerosal-autism connection is still missing the
“Thimerosal-controversy” has stimulated the filing of a number of lawsuits in the US by
affected individuals that seek compensation from vaccine producing companies. But given the
rather limited knowledgebase on the toxicity, distribution and metabolism of Thimerosal in
higher organisms it certainly seems difficult to conclude whether or not Thimerosal
containing vaccines might cause neuro developmental disorders. Systematic research studies
aimed at these topics are therefore needed, as was concluded by the IOM in their 2001 report
[10]. The toxicological effects of ethylmercury, which is the toxic moiety of Thimerosal, are
understudied and most often researchers have assumed its toxicity to be analogous to that of
methylmercury. It has been suggested that ethylmercury is rapidly released from the
Thimerosal moiety upon its administration in the human body and therefore disposition
pathways similar to those of methylmercury can be expected [4]. However, ethylmercury
seems to be less stable than methylmercury within the body, which may result in toxilogical
differences between the two compounds. For example, Matheson et al [100] reported five fold
higher blood levels of inorganic mercury for a patient exposed to large amounts of Thimerosal
compared to what is usually found in cases of methylmercury poisoning of similar degree.
In paper III a multi-labelled SSID GC-ICP-MS method was developed to study the
disposition and metabolism/transformation of Thimerosal in mice. As mentioned the
ethylmercury moiety of Thimerosal is rapidly dissociated from the thiosalicylic acid upon its
23
introduction in biological systems and therefore an ethylmercury isotopomer was synthesised
from enriched 199HgO to enable SSID quantitation of this metabolite. By using the
comproportionation reaction previously applied for the synthesis of methylmercury
isotopomers in paper I, C2H5199HgCl could be produced with a yield of >90 %. The isotope
labelled ethylmercury chloride was subsequently transferred from the reaction medium
(toluene) to aqueous media by evaporating the organic solvent and redissolving the resulting
dry white powder in 10 % aqueous ethanol solution. In this way the isotopomer could be
added to samples in a medium that was more compatible with the biological tissues under
study and, in addition, unwanted residual mercury by-products like diethylmercury were
removed to a large extent facilitated by the moderate heat evaporation of the organic solvent.
In addition to the ethylmercury isotope spike the samples were also standardised with
CH3200HgCl and 201Hg2+. Accurate results for the methyl- and inorganic mercury content of
two certified reference materials (CRM) were obtained. Ethylmercury was added to the
CRMs in varying concentrations and quantitative recoveries were obtained. Nevertheless, it
was found from the isotope ratio measurements that almost one tenth of the ethylmercury
isotopomer, and consequently also the incipient ethylmercury present in the sample, was
degraded to inorganic mercury during sample pre-treatment and analysis. Given that such
species conversions are very difficult to correct for when using non-isotope selective
techniques, it could be concluded that SSID calibration is essential for achieving accurate
results in ethylmercury determinations.
Species concentrations of methyl-, ethyl- and inorganic mercury were assessed in
duplicates in kidney, liver and mesenterial lymph nodes following 0 (control), 1, 2.5, 6 and 14
days of Thimerosal treatment ad libitum (figure 10). A rapid absorption of the ethylmercury
moiety of Thimerosal in the experimental animals was reflected in the ethylmercury
concentrations observed in the analysed mouse tissues. However, it was clearly evident that
ethylmercury displayed limited stability once absorbed as found concentrations of inorganic
mercury in the organs increased with time, which was in line with the observations made by
Matheson et al. Interestingly, methylmercury concentrations also increased with time in the
experimental animals, but this observation could be linked to methylmercury impurities in the
mouse food as well as the administered Thimerosal.
24
0.8
60
0.6
40
0.4
20
0.2
0
0.0
+
80
CH3Hg conc. / µg g-1
1.0
+
C2H5Hg and Hg2+ conc. / µg g-1
a
100
0
2
4
6
8
10
12
14
b
0.30
0.25
20
0.20
15
0.15
10
0.10
5
0.05
0
0.00
+
25
CH3Hg conc. / µg g-1
30
+
C2H5Hg and Hg2+ conc. / µg g-1
Days of Thimerosal exposure
0
2
4
6
8
10
12
14
8
c
0.08
0.06
4
0.04
2
0.02
0
0.00
+
6
CH3Hg conc. / µg g-1
+
C2H5Hg and Hg2+ conc. / µg g-1
Days of Thimerosal exposure
0
2
4
6
8
10
12
14
Days of Thimerosal exposure
Figure 10. Determined (+) CH3Hg+, ( ) C2H5Hg+ and ( ) Hg2+concentrations in a) kidney b) liver
and c) mesenterial lymph nodes as a function of days of Thimerosal exposure.
25
4.3
Towards a wider implementation of SSID methodology in mercury
speciation, papers IV and V
Species-specific isotope dilution facilitates considerably simpler and more flexible analytical
protocols in terms of sample processing compared to most other methods for mercury
speciation analysis. So far though, the methodology has mainly been used as a reference
technique by a relatively few research laboratories. A wider implementation on routine basis
at commercial laboratories has been hindered by the lack of available certified isotope
standards and to some extent lack of certified reference materials [101], which are a requisite
for quality assurance in terms of accuracy and traceability of results at accredited analytical
laboratories.
Paper IV describes the preparation and certification of aqueous methylmercury
isotopic certified reference material (ICRM) enriched to 98 % in the 202Hg isotope with a
certified methylmercury amount content and isotopic composition that is traceable to SI-units.
This ICRM will hopefully help launch SSID calibration as a standard method in mercury
speciation analysis. The synthesis route that had previously been used successfully for the
production of methyl- and ethylmercury isotopomers [I, III] was again applied here, although
in a somewhat larger scale. However, special emphasis was in this case put into producing as
pure a product as possible to minimise the risk of unwanted side reactions that potentially
could influence the species integrity. This meant applying product purification procedures to
quantitatively remove residual mercury species in the raw product, e.g. Hg2+, (CH3)2Hg and
Hg0. Two independent yearlong stability tests were carried out to investigate the stability of
the material. One conventional, in which the standard was stored at -20 ° C and the
methylmercury amount content analysed every three months, and one isochronous stability
test, where six sets of standard aliquots were stored at temperatures of -20 ° C, 4 ° C and 20 °
C and were then analysed all at the same occasion. The results clearly showed that the
isotopomer standard was stable for at least one year independent of storage temperature. The
resulting aqueous CH3202Hg standard was certified with total CH3Hg as well as CH3202Hg
amount contents of 35.3 and 34.5 µg g-1, respectively. Inorganic mercury, which was the only
detectable residual species, was found to be present at 2 % of the total mercury content. The
CH3202Hg ICRM, named ERM-AE670, is currently commercially available through European
Reference Materials, ERM [102].
A potential application for the ICRM developed in paper IV is for routine
determinations of methylmercury in water samples. Natural waters are perhaps the most
difficult samples to deal with in mercury speciation analysis owing to the very low levels of
mercury species present and their limited stability during sample storage, methylmercury in
particular (section 3.1). Analyte instability should be efficiently corrected for if speciesspecific standardisation is applied during sampling since losses of analyte and added
isotopomer will occur to equal extents, given that the isotopomer has been properly
equilibrated with the sample. In that way, previous requirements associated with non-SSID
methods of keeping sample storage time as short as possible to prevent erroneous results from
analyte degradation/losses could be evaded. This would be very convenient during e.g. longterm studies since samples then could be accumulated batch wise and analysed accordingly at
a suitable occasion instead of having to analyse samples continuously. Likewise, by applying
SSID standardisation to water samples it should also be possible circumvent the need for
analyte-matrix separation prior to derivatisation that is required with non-SSID methods
(section 3.1) since non-quantitative analyte recoveries due to matrix related interferences will
not effect the accuracy of results. Thereby sample preparation protocols could be kept
extremely simple. The above hypotheses were tested in paper V in which a field-adapted
26
method for methylmercury determination in natural water samples based on SSID
standardisation was developed and applied in the field. The possibility of direct analyte
derivatisation through aqueous phase ethylation using sodium tetraethylborate without any
kind of sample treatment (other than filtration of suspended particles) was tested against
conventional liquid-liquid (quantitative) extractions of CH3200HgCl standardised water
samples of varying matrix composition with respect to salinity and DOC content (mire-,
fresh- and seawater). Nearly identical results were obtained between the two methods and it
could thus be concluded that SSID based methylmercury determinations can be conducted in
a wide range of natural waters without any analyte-matrix separation prior to analyte
derivatisation. Furthermore, it was observed that the applied liquid-liquid extraction treatment
generated methylmercury blank levels, and accordingly detection limits, an order of
magnitude higher than the direct ethylation approach. Although the application of SSID
dramatically simplified sample handling and treatment protocols, it should be pointed out that
analyte loss at trace level concentrations could turn out to be crucial in terms of detectability.
It was thus desirable to combine SSID sample standardisation with some kind of sample
preservation procedure to subdue analyte losses. Therefore sample storage at –20 º C was
investigated and results were compared to those obtained from analysing the same samples as
soon as possible after sampling. Again, no significant differences in either analyte
concentrations or recoveries between freeze stored or immediately analysed samples could be
detected in the tested matrices, which showed that samples could be safely stored in a regular
household freezer.
Adapting the developed method for the field study meant that samples had to be
standardised and frozen in conjunction to sampling. For simplicity, the CH3200HgCl
isotopomer was added volumetrically by a micropipette even though this procedure was likely
to have a greater detrimental effect on the overall uncertainty of the measurements than if the
standard had been added gravimetrically. However, total uncertainty budgets for the field
application of the method revealed that the highest uncertainty contributions derived from,
depending on the analyte concentration, either uncertainty of the concentration of the
isotopomer standard or from variation in field blanks. Nevertheless, the relative expanded
uncertainties derived from the uncertainty budgeting ranged between 5-13 % for the observed
methylmercury concentration range in 50 analysed mire water samples, which was sufficient
to allow for detection of small differences in analyte concentration at trace level.
5. Conclusions and prospects
The work presented in this thesis demonstrates the potential and usefulness of SSID
methodology in mercury speciation, either if it is applied solely for quantification or if it is
combined with isotope enriched tracers for studies of mercury species transformations in
marine, or other, environments. In the latter case, benefits can be gained with respect to the
number of simultaneously observable parameters, as well as the specificity and precision at
which they can be measured. The method presented in paper I and similar SSID based
approaches presented by others can, in the authors opinion, be considered as more or less
perfect for the purpose of assessing mercury species transformation rates in sediments. With
this methodology available the conditions to further resolve biological and geochemical
mechanisms that control microbial mercury methylation are quite favourable. However,
information regarding tracer versus incipient species availability for species transformations
and its effect on observed transformation rates in sediments is scarce [85] and more research
is needed on this topic. This knowledge would facilitate finer tuned stable isotope tracer
27
methods. In wider perspective, it is important to include other environmental compartments of
the “mercury cycle” in mercury species transformation studies e.g. water, biota and
atmosphere. In that way it should be possible to obtain good quality data on important
parameters such as mercury species biomagnification and “inter-compartment” transfer rates.
Such experiments, based on microcosm simulations of brackish water systems, are currently
under way in our laboratory utilizing newly developed SSID-based technology for
measurement of gaseous mercury species [103].
The developed SSID method for determination of ethylmercury evidently enables
accurate determinations of this apparently unstable mercury species in biological samples.
The next logical step to further develop the method should involve synthesis of isotope
enriched Thimerosal that is administered to experimental animals followed by SSID
calibration using an ethylmercury isotopomer labelled in a different isotope in the analytical
stage. This would certainly make a “sharper” analytical tool at hand for Thimerosal toxicity
studies in the way that traceability of in-situ species conversions would be much improved.
Paper V showed that SSID makes sample storage less critical and reduces sampling
treatment procedures associated with natural water samples. At the same time, accuracy of the
analytical results is maintained. These features should certainly appeal to commercial
laboratories that handle large quantities samples. However, accurate in-field SSID
standardisation of water samples (or other sample matrices) is still hampered by the
requirement of background data on analyte concentration for the sample under study. A
solution to this problem could be to standardise samples with an assortment of different
methylmercury isotopomers that cover a wide concentration range. In that way the probability
to obtain an optimal isotopomer/reference ratio, to minimise measurement uncertainty for an
unknown sample, is increased.
Finally, from an analytical viewpoint it is desirable to develop mercury speciation
methods in analogy to the IUPAC definition of speciation as being “determination of the
exact chemical form or compound in which an element occurs in a sample” (see introduction).
With respect to this, the GC-ICP-MS technique commonly used in mercury speciation, and
throughout this thesis, can fulfil these criteria for elemental mercury but is unfortunately a
blunt technique when other mercury species are concerned. Future efforts should therefore be
aimed at developing complementary methods based on molecular mass spectrometric
techniques, such as LC-MS(-MS) or GC-MS, as well as improvement of existing direct
methods.
28
6. Acknowledgements
Wolfgang: Först och främst ska du ha tack för alla ovärderliga synpunkter på laborativa och
skriftliga moment samt att du i de allra flesta fall ser till att allt “snack” resulterar i nånting
konkret. Stort tack även för alla festliga sammankomster du arrangerar-höjdpunkter under min
tid som doktorand.
Erik L: Tack för alla tips och råd och för att du alltid förbehållslöst ställt upp och ordnat både
det ena och det andra.
Mats: Tack för din entusiasm och alla högintressanta projekt (gamla som nya). Dina
kunskaper och idéer har haft avsevärd betydelse.
Ett jättetack går ut till alla fantastiska medarbetare/vänner i AAS-gruppen, nuvarande och
föredettingar, som på ett eller annat sett bidragit till att göra de här åren så roliga och
stimulerande. Det har varit ett nöje att arbeta tillsammans, tack: Tom för att du låter mej ta
hand om tjädrarna i Vargträsk och andra ställen-absoluta höjdpunkter! Ser fram emot 25/8!,
Erik B för givande diskussioner, “laborationer” och kollaborationer rörande diverse
vattenlösningar samt för alla klarsynta kommentarer, Johanna för givande samarbete, lycka
till med familj och karriär, Solomon för din positiva attityd och stöd genom åren, James för
all teknisk hjälp i början och för gemensamma projekt på senare tid, Dong and Daniel - good
luck in the future!, Karin, Jin och Qiang.
Tack också till övriga kollegor på analytisk kemi: Lasse för all hjälp med instrument, datorer
och trevliga jaktturer på ditt natursköna hemman, Svante J för tekniska konstruktioner, Patrik,
Tobias o Nina & Co., Petrus, William, Michael, Svante Å, Camilla, Anna S, Anders, Josefina,
Knut, Anna N, Julien, Britta, Wen, Christer, Emil, Ahn Mai, Fredrik, Erika, Einar, Jufang och
Gerd.
De senaste året har jag varit relativt osynlig på min andra arbetsplats UMF av olika skäl
(boken ni läser just nu t ex). De perioder som jag tillbringat på UMF har emellertid varit
otroligt trevliga, vilket till stor del berott på personalen-tack allihop! Särskilt tack till Calle
och Tommy för laborativ assistans samt Mikael och Jonas för assistans i jakten på
svavelosande gegga.
Tack även till SLU:are Ulf, Andreas, Torbjörn och Kevin.
Per Hultman och Said Havarinasab i Linköping ska ha stort tack för ett givande samarbete. Vi
kanske ses IRL nångång!
Tack till klanmedlemmar som på ett eller annat sätt bidragit till att underlätta doktorandlivet:
Morsan, Einar, Karin, Jörgen, Kalle, Elsa, Edvin, Farsan, Eva, Oskar, Lisa, Ulla, Börje och
Kicki.
Slutligen vill jag ge min underbara familj ett gäng jättekramar: Lotta, Max, Greta och
Elvira, NI ÄR BÄST!
29
7. References
1.
2.
3.
4.
5.
6.
IUPAC Compendium of Chemical Terminology, 2nd Edition, 1997.
European Water Framework Directive, 2000, 60, EC.
Global Mercury Assessment, UNEP Chemicals, Geneva, Switzerland, 2002.
Clarkson, T. W., Environ. Health Persp., 110, 11-23 (2002)
Eto, K., Neuropathology, 2000, 20: S14.
Bernard, S., Enayati, A., Redwood, L., Roger, H., Binstock, T., Med. Hypotheses,
2001, 56, 462-471.
7. Langford, N. J., Ferner, R. E., J. Hum. Hypertens., 1999, 13, 651-656.
8. Harada, M., Crit. Rev. Toxicol., 1995, 25, 1.
9. Methylmercury, Environmental Health Criteria 101, 1991, World Health
Organization, Geneva.
10. Immunization Safety Review Committee: Thimerosal-Containing Vaccines and
Neurodevelopmental Disorders. 2001, US Institute of Medicine.
11. Swedish Expert Group, Methylmercury in fish, A toxicological epidemiological
evaluation of risks, Nord Hyg Tidskr 4(suppl), 1971, 19.
12. Håkanson, L., Nilsson, Å., Andersson, T., Water Air Soil Poll., 1988, 50, 171.
13. Westöö. G., Acta Chem. Scand., 1966, 20, 2131-2137.
14. Jensen, S., Jernelöv, A., Nature, 1969, 223, 753.
15. Joint FAO/WHO Expert Committee on Food Additives, 61st meeting, Rome 10-19
June 2003.
16. Lamborg, C. H., Fitzgerald, W. F., O’Donnel, J., Torgersen, T., Geochim.
Cosmochim. Acta., 2002, 66, 1105-1118.
17. Mercury Study Report to Congress, US EPA, Dec. 1997.
18. Compeau, G. C., Bartha, R., Appl. Environ. Microbiol., 1985, 50, 498.
19. Choi, S., Bartha, R., Appl. Environ. Microbiol., 1993, 59, 290.
20. Choi, S., Chase, T., Bartha, R., Appl. Environ. Microbiol., 1994, 60, 1342.
21. Ekstrom, E. B., Morel, M. M. F., Benoit. J. M., Appl. Environ. Microbiol., 2003, 69,
5414.
22. Landner, L., Nature, 1971, 230, 452.
23. Devereux, R., Winfrey, M. R., Winfrey, J., Stahl, D. A., FEMS Microb. Ecol., 1996,
20, 23.
24. King, J. K., Saunders, F. M., Lee, R. F., Jahnke, R. A., Environ. Toxicol. Chem., 1999,
18, 1362.
25. King, J. K., Kostka, J. E., Frischer, M. E., Saunders, F. M., Appl. Environ. Microbiol.,
2000, 66, 2430.
26. King, J. K., Kostka, J. E., Frischer, M. E., Saunders, F. M., Jahnke, R. A., Environ.
Sci. Technol., 2001, 35, 2491.
27. Baldi, F., Parati, F., Filippeli, M., Water Air Soil Pollut., 1995, 80, 805.
28. Benoit, J. M., Gilmour, C. C., Mason, R. P., Appl. Environ. Microbiol., 2001, 67, 51.
29. Baldi, F., Pepi, M., Filippelli, M., Appl. Environ. Microbiol., 1993, 59, 2479.
30. Robinson, J. B., Tuovinen, O. L., Microbiol. Rev., 1984, 48, 95.
31. Rugh, C. L., Senecoff, J. F., Meagher, R. B., Merkle, S. A., Nature Biotechnol. 1998,
16, 925.
32. Marvin-DiPasquale, M. C., Oremland, R. S., Environ. Sci. Technol., 1998, 32, 2556.
33. Marvin-DiPasquale, M., Agee, J., McGowan, C., Oremland, R. S., Thomas, M.,
Krabbenhoft, D., Gilmour, C. C., Environ, Sci. Technol., 2000, 34, 4908.
30
34. Benoit, J. M., Gilmour, C. C., Heyes, A., Mason, R. P., Miller, C. L., Biogeochemistry
of Environmentally Important Trace Elements, eds Cai, Y., Braids, O. C., 2003,
American Chemical Society, Washington, DC.
35. Xia, K., Skyllberg, U., Bleam, W. F., Bloom, P. R., Nater, E. A., Helmke, P. A.,
Environ. Sci. Technol., 1999, 33, 257.
36. Qian, J., Skyllberg, U., Frech, W., Bleam, W. F., Bloom, P. R., Petit, P. E., Geochim.
Cosmochim. Acta, 2002, 66, 3873.
37. Method 1669, “Method for Sampling Ambient Water for Determination of Metals at
EPA Ambient Criteria Levels”, U. S. Environment Protection Agency, Washington
DC. 1996.
38. Reinholdsson, F., Briche, C., Emteborg, H., Baxter, D. C., Frech, W., Proceedings
following CANAS ´95.
39. Bloom, N. S., Can. J. Fish. Aquat. Sci., 1989, 46, 1131.
40. Achmed, R., Stoeppler, M., Anal. Chim. Acta, 1987, 192, 109
41. Achmed, R., Hay, K., Stoeppler, M., Fresenius J. Anal. Chem., 1987, 326, 510.
42. Johansson, M., Emteborg, H., Glad, B., Reinholdsson, F., Baxter, D. C., Fresenius J.
Anal. Chem., 1995, 351, 461.
43. Crecelius, E. A., Bloom, N. S., Cowan, C. E., Jenne, E. A., Speciation of selenium and
arsenic in natural waters and sediments, Vol. 2. EPRI EA-4641. Electric Power
Research Institute, Palo Alto, CA.
44. Horvat, M., Liang, L., Bloom, N. S., Anal. Chim. Acta, 1993, 282, 153.
45. Bowles, K. C., Apte, S. C., Anal. Chem., 1998, 70, 395.
46. Emteborg, H., Baxter, D. C., Frech, W., Analyst, 1993, 118, 1007.
47. Yamini, Y., Alizadeh, N., Shamsipur, M., Anal. Chim. Acta, 1997, 355, 69.
48. Cai, Y., Bayona, J. M., J. Chromatogr. A., 1995, 696, 113.
49. Cai. Y., Monsalud, S., Furton, K. G., Jaffe, R., Jones, R. D., Appl. Organomet. Chem,
1998, 12, 565.
50. Saouter, E., Blattmann, B., Anal. Chem., 1994, 66, 2031.
51. Qvarnström, J., Tu, Q., Frech, W., Lüdke, C., Analyst, 2000, 125, 1193.
52. Demuth, N., Heumann, K. G., Anal. Chem., 2001, 73, 4020.
53. Westöö, G., Acta Chem. Scand., 1966, 20, 2131.
54. Huang J. H., Ilgen G., Matzner E., Anal. Chim. Acta, 2003, 493, 23.
55. Emteborg, H., Björklund, E., Ödman, F., Karlsson, L., Mathiasson, L., Frech, W.,
Baxter, D. C., Analyst, 1996, 121, 19.
56. Bayona, J. M., Trend. Anal. Chem., 2000, 19, 107.
57. Tseng, C. M., DeDiego, A., Martin, F. M., Amouroux, D., Donard, O. F. X., J. Anal.
At. Spectrom., 1997, 12, 743.
58. Kim, A. W., Foulkes, M. E., Ebdon, L., Hill, S. J., Patience R. L., Barwise, A. G.,
Rowland, S. J., J. Anal. At. Spectrom., 1992, 7, 1147.
59. Heumann, K. G., in: Broekaert, J. A. C., Güçer, S., Adams, F. (Eds), Metal Speciation
in the Environment, NATO ASI Series, Vol. G 23, Springer Verlag, Heidelberg,
Germany, 1990, p. 153.
60. Heumann, K. G., Int. J. Mass Spectrom. Ion Processes, 1992, 118/119, 575.
61. Hoezl, R., Hoezl, C., Kotz, L., Fabry, L., Accred. Qual. Assur., 1998, 3, 185.
62. Furata, N., Ohata, M., Ichinose, T., Shinohara, A., Chiba, M., Anal. Chem., 1998, 70,
2726.
63. Rouleau, C., Block, M., Appl. Organomet. Chem., 1997, 11, 751.
64. Cappon, C. J., Smith, J. C., Anal. Chem., 1977, 49, 365.
65. Hintelmann, H., Evans, R. D., Fresenius J. Anal. Chem., 1997, 358, 378.
31
66. Rahman, G. M. M., Kingston, H. M. S., Bhandari, S., Appl. Organomet. Chem., 2003,
17, 913.
67. Snell, J. P., Stewart, I., Sturgeon, R. E., Frech, W., J. Anal. At. Spectrom., 2000, 15,
1540.
68. Siegler, R. W., Nierenberg, D. W., Hickey, W. F., Hum. Pathol., 1999, 30, 720.
69. Nusko, R., Heumann, K. G,. Anal. Chim. Acta, 1994, 286, 283.
70. Ruiz Encinar, J., Garcia Alonso, J. I., Sanz-Medel, A., J. Anal. At. Spectrom., 2000,
15, 1233.
71. Kingston, H. M., Huo, D., Lu, Y., Chalk, S., Spectrochim. Acta B, 1998, 53, 299.
72. Ruiz Encinar, J., Rodriguez-González, P., Garcia Alonso, J. I., Sanz-Medel, A., Anal.
Chem., 2002, 74, 270.
73. Qvarnström, J., Frech, W., J. Anal. At. Spectrom., 2002, 17, 1486.
74. Hintelmann, H., Falter, R., Ilgen, G., Evans, R. D., Fresenius J. Anal. Chem., 1997,
358, 363.
75. Mason, R. P., Fitzgerald, W. F., Hurley, J., Donaghay, P. L., Seiburth, L. P., Limnol.
Oceanogr., 1992, 38, 1227.
76. Hintelmann, H., Chemosphere, 1999, 39, 1093.
77. Falter, R., Chemosphere, 1999, 39, 1051.
78. Falter, R., Chemosphere, 1999, 39, 1075.
79. Rodriguez Martin-Doimeadios. R. C., Monperrus, M., Krupp, E., Amouroux, D.,
Donard, O. F. X., Anal. Chem., 2003, 75, 3202.
80. Furatani, A., Rudd, J. W. M., Appl. Environ. Microbiol., 1980, 40, 770.
81. Ramlal, P. S., Rudd, J. W. M., Hecky, R. E., Appl. Environ. Microbiol., 1986, 51, 110.
82. Gilmour, C. C., Riedel, G. S., Water Air Soil Pollut., 1995, 80, 747.
83. Hintelmann, H., Evans, R. D., Villeneuve, J. Y., J. Anal. Atom. Spectrom., 1995, 10,
619.
84. Hintelmann, H., Evans, R. D., Fresenius J. Anal. Chem., 1997, 358, 378.
85. Hintelmann, H., Keppel-Jones, K., Evans, R. D., Environ. Toxicol. Chem. 2000, 19,
2204.
86. Taylor, H. E., Huff, R. A., Montaser, A., In Inductively Coupled Plasma Mass
Spectrometry, Montaser, A., Ed., WILEY-VHC, 1998, pp 681-807.
87. Wehmeier, S., Ellam, R., Feldmann, J., J. Anal. At. Spectrom., 2003, 18, 1001.
88. Korthals, E. T. Winfrey, M. R., Appl. Environ. Microbiol., 1987, 53, 2397.
89. Hines, M. E., Horvat, M., Faganeli, J., Bonzongo, J. C. J., Barkay, T., Major, E. B.,
Scott, K. J., Bailey, E. A., Warwick, J. J., Lyons, W. B., Environ. Res., 2000, 83, 129.
90. Devereux, R., Winfrey, M. R., Winfrey, J., Stahl, D. A., FEMS Microbiol. Ecol.,
1996, 20, 23.
91. Barkay, T., Gillman, M., Turner, R. R., Appl. Environ. Microbiol., 1997, 63, 4267.
92. Gilmour, C.C., Riedel, G. S., Ederington, M. C., Bell, J. T., Benoit, J. M., Gill, G.A.,
Stordal, M. C., Biogeochem., 1998, 40, 327.
93. Benoit, J. M., Gilmour, C. C., Mason, R. P., Heyes, A., Environ. Sci Technol., 1999,
33, 951.
94. Benoit, J. M., Gilmour, C. C., Mason, R. P., Environ. Sci. Technol., 2001, 35, 127.
95. Ramlal, P. S., Kelly, C. A., Rudd, J. W. M., Can. J. Fish. Aquat. Sci., 1992, 50, 972.
96. Hammerschmidt, C. R., Fitzgerald, W. F., Environ. Sci. Technol., 2004, 38, 1487.
97. Kwokal, Z., Franciskovic-Bilinski. S., Bilinski, H., Branica, M., Mar. Pollut. Bull.,
2002, 44, 1152.
98. Geier, M. R., Geier, D. A., Exp. Biol. Med., 2003, 228, 660.
99. Nelson, K. B., Bauman, M. L., Pediatrics, 2003, 111, 674.
100. Matheson, D. S., Clarkson, T. W., Gelland, E. W., J. Pediatr. 1980, 97, 153.
32
101. Heumann, K. G., Anal. Bioanal. Chem., 2004, 378, 318.
102. http://www.erm-crm.org/ermcrm
103. Larsson, T., Frech, W., Anal. Chem., 2003, 75, 5584.
33