Phytoremediation of mercury by terrestrial plants Yaodong Wang

Phytoremediation of mercury
by terrestrial plants
Yaodong Wang
Department of Botany Stockholm University, 2004
Phytoremediation of mercury by terrestrial plants
A doctoral dissertation to be public defended 21 December 2004 at 10:00 in the lecture hall
at the Department of Botany, Stockholm University
Faculty opponent: Professor Douglas Godbold, School of Agricultural & Forest Sciences,
University of Wales, UK.
Mercury (Hg) pollution is a global environmental problem. Numerous Hg-contaminated sites exist in the
world and new techniques for remediation are urgently needed. Phytoremediation, use of plants to remove
pollutants from the environment or to render them harmless, is considered as an environment-friendly
method to remediate contaminated soil in-situ and has been applied for some other heavy metals. Whether
this approach is suitable for remediation of Hg-contaminated soil is, however, an open question. The aim
of this thesis was to study the fate of Hg in terrestrial plants (particularly the high biomass producing
willow, Salix spp.) and thus to clarify the potential use of plants to remediate Hg-contaminated soils.
Plants used for phytoremediation of Hg must tolerate Hg. A large variation (up to 30-fold difference)
was detected among the six investigated clones of willow in their sensitivity to Hg as reflected in their
empirical toxicity threshold, the maximum unit toxicity and EC50 levels. This gives us a possibility to
select Hg-tolerant willow clones to successfully grow in Hg-contaminated soils for phytoremediation.
Release of Hg into air by plants is a concern when using phytoremediation in practice. No evidence
was found in this study that Hg was released to the air via shoots of willow, garden pea (Pisum sativum
L.), spring wheat (Triticum aestivum L.), sugar beet (Beta vulgaris L.), oil-seed rape (Brassica napus L.)
and white clover (Trifolium repens L.). Thus, we conclude that the Hg burden to the atmosphere via
phytoremediation is not increased.
Phytoremediation processes are based on the ability of plant roots to accumulate Hg and to
translocate it to the shoots. Willow roots were shown to be able to efficiently accumulate Hg in
hydroponics, however, no variation in the ability to accumulate was found among the eight willow clones
using CVAAS to analyze Hg content in plants. The majority of the Hg accumulated remained in the roots
and only 0.5-0.6% of the Hg accumulation was translocated to the shoots. Similar results were found for
the five common cultivated plant species mentioned above. Moreover, the accumulation of Hg in willow
was higher when being cultivated in methyl-Hg solution than in inorganic Hg solution, whereas the
translocation of Hg to the shoots did not differ.
The low bioavailability of Hg in contaminated soil is a restricting factor for the phytoextraction of
Hg. A selected tolerant willow clone was used to study whether iodide addition could increase the plantaccumulation of Hg from contaminated soil. Both pot tests and field trials were carried out. Potassium
iodide (KI) addition was found to mobilize Hg in contaminated soil and thus increase the bioavailability
of Hg in soils. Addition of KI (0.2–1 mM) increased the Hg concentrations up to about 5, 3 and 8 times in
the leaves, branches and roots, respectively. However, too high concentrations of KI were toxic to plants.
As the majority of the Hg accumulated in the roots, it might be unrealistic to use willow for
phytoextraction of Hg in practice, even though iodide could enhance the phytoextraction efficiency.
In order to study the effect of willow on various soil fractions of Hg-contaminated soil, a 5-step
sequential soil extraction method was used. Both the largest Hg-contaminated fractions, i.e. the Hg bound
to residual organic matter (53%) and sulphides (43%), and the residual fraction (2.5%), were found to
remain stable during cultivations of willow. The exchangeable Hg (0.1%) and the Hg bound to humic and
fulvic acids (1.1%) decreased in the rhizospheric soil, whereas the plant accumulation of Hg increased
with the cultivation time. The sum of the decrease of the two Hg fractions in soils was approximately
equal to the amount of the Hg accumulated in plants. Consequently, plants may be suitable for
phytostabilization of aged Hg-contaminated soil, in which root systems trap the bioavailable Hg and
reduce the leakage of Hg from contaminated soils.
Yaodong Wang
Department of Botany
Stockholm University, Sweden
© Yaodong Wang 2004
ISBN 91-7265-975-0 pp 1–41
[email protected]
Phytoremediation of mercury
by terrestrial plants
Yaodong Wang
王 耀 东
Doctoral thesis
Department of Botany
Stockholm University, 2004
1
Doctoral dissertation
Yaodong Wang
Department of botany
Stockholm University
S-106 91 Stockholm
[email protected]
December 21, 2004
Front cover: Illustration of phytostabilization of Hg-contaminated soil
© Yaodong Wang
ISBN 91-7265-975-0 pp.1–41
Printed at PrintCenter, Stockholm University
Stockholm, Sweden, 2004
2
Abstract
Mercury (Hg) pollution is a global environmental problem. Numerous Hg-contaminated sites exist
in the world and new techniques for remediation are urgently needed. Phytoremediation, use of
plants to remove pollutants from the environment or to render them harmless, is considered as an
environment-friendly method to remediate contaminated soil in-situ and has been applied for some
other heavy metals. Whether this approach is suitable for remediation of Hg-contaminated soil is,
however, an open question. The aim of this thesis was to study the fate of Hg in terrestrial plants
(particularly the high biomass producing willow, Salix spp.) and thus to clarify the potential use of
plants to remediate Hg-contaminated soils.
Plants used for phytoremediation of Hg must tolerate Hg. A large variation (up to 30-fold
difference) was detected among the six investigated clones of willow in their sensitivity to Hg as
reflected in their empirical toxicity threshold (TT95b), the maximum unit toxicity (UTmax) and EC50
levels. This gives us a possibility to select Hg-tolerant willow clones to successfully grow in Hgcontaminated soils for phytoremediation.
Release of Hg into air by plants is a concern when using phytoremediation in practice. No
evidence was found in this study that Hg was released to the air via shoots of willow, garden pea
(Pisum sativum L. cv Faenomen), spring wheat (Triticum aestivum L. cv Dragon), sugar beet (Beta vulgaris
L. cv Monohill), oil-seed rape (Brassica napus L. cv Paroll) and white clover (Trifolium repens L.). Thus,
we conclude that the Hg burden to the atmosphere via phytoremediation is not increased.
Phytoremediation processes are based on the ability of plant roots to accumulate Hg and to
translocate it to the shoots. Willow roots were shown to be able to efficiently accumulate Hg in
hydroponics, however, no variation in the ability to accumulate was found among the eight willow
clones using CVAAS to analyze Hg content in plants. The majority of the Hg accumulated remained
in the roots and only 0.5-0.6% of the Hg accumulation was translocated to the shoots. Similar
results were found for the five common cultivated plant species mentioned above. Moreover, the
accumulation of Hg in willow was higher when being cultivated in methyl-Hg solution than in
inorganic Hg solution, whereas the translocation of Hg to the shoots did not differ.
The low bioavailability of Hg in contaminated soil is a restricting factor for the phytoextraction
of Hg. A selected tolerant willow clone was used to study whether iodide addition could increase
the plant-accumulation of Hg from contaminated soil. Both pot tests and field trials were carried
out. Potassium iodide (KI) addition was found to mobilize Hg in contaminated soil and thus increase
the bioavailability of Hg in soils. Addition of KI (0.2–1 mM) increased the Hg concentrations up to
about 5, 3 and 8 times in the leaves, branches and roots, respectively. However, too high
concentrations of KI were toxic to plants. As the majority of the Hg accumulated in the roots, it
might be unrealistic to use willow for phytoextraction of Hg in practice, even though iodide could
enhance the phytoextraction efficiency.
In order to study the effect of willow on various soil fractions of Hg-contaminated soil, a 5-step
sequential soil extraction method was used. Both the largest Hg-contaminated fractions, i.e. the Hg
bound to residual organic matter (53%) and sulphides (43%), and the residual fraction (2.5%), were
found to remain stable during cultivations of willow. The exchangeable Hg (0.1%) and the Hg
bound to humic and fulvic acids (1.1%) decreased in the rhizospheric soil, whereas the plant
accumulation of Hg increased with the cultivation time. The sum of the decrease of the two Hg
fractions in soils was approximately equal to the amount of the Hg accumulated in plants.
Consequently, plants may be suitable for phytostabilization of aged Hg-contaminated soil, in which
root systems trap the bioavailable Hg and reduce the leakage of Hg from contaminated soils.
3
List of papers
The present thesis is based on the following papers, which will be referred to by their
Roman numerals.
I.
Greger, M., Wang, Y.D.*, Neuschütz C., 2005. Absence of Hg transpiration by
shoot after Hg uptake by roots of six terrestrial plant species. Environmental
Pollution (in press).
II.
Wang, Y.D.*, Greger, M., 2004. Clonal differences in mercury tolerance,
accumulation and distribution in Willow. Journal of Environmental Quality
33:1779-1785.
III.
Wang, Y.D.*, Stauffer, C., Keller, C., Greger, M. Changes in Hg fractionation in
soil induced by willow. Plant and Soil (accepted)
IV.
Wang, Y.D.*, Greger, M. Use of iodide to enhance the phytoextraction of mercurycontaminated Soil. (Submitted to Science of the Total Environment)
Reprints and accepted papers are published by kind permission of the journals concerned.
* Corresponding author
4
Table of Contents
Abstract................................................................................................................................ 3
List of papers ....................................................................................................................... 4
Table of Contents................................................................................................................. 5
Abbreviations....................................................................................................................... 6
Introduction.......................................................................................................................... 7
Hg - a global environmental pollutant......................................................................... 7
Sources of Hg pollutants .................................................................................. 7
Hg speciation in air, water and soil .................................................................. 8
Health risks of Hg........................................................................................... 10
Overview of Phytoremediation ................................................................................. 10
Advantages and disadvantages of phytoremediation...................................... 11
Processes and technologies of phytoremediation of heavy metals ................. 11
Plant interaction with Hg........................................................................................... 12
Aim of the present study .................................................................................................... 14
Comments on materials and methods ................................................................................ 15
Plant materials ........................................................................................................... 15
Plant chamber systems and Hg traps......................................................................... 15
Weibull function........................................................................................................ 17
Hg analyses ............................................................................................................... 17
Results and discussions...................................................................................................... 19
Hg accumulation and distribution in six terrestrial plant species .............................. 19
Hg accumulation and distribution in eight willow clones ......................................... 19
Accumulation and distribution of various Hg species in willow............................... 20
Screening of Hg accumulators in a Hg-contaminated site......................................... 24
Hg exchange between plants and air ......................................................................... 26
Sensitivity of willow to Hg ....................................................................................... 28
Phytoremediation of Hg ............................................................................................ 30
Phytoextraction............................................................................................... 30
Phytostabilization ........................................................................................... 32
Foliage filtration ............................................................................................. 32
Conclusions........................................................................................................................ 34
Future perspectives ............................................................................................................ 34
Acknowledgements............................................................................................................ 35
References.......................................................................................................................... 37
5
Abbreviations
AFS: Atomic fluorescence spectrometry
+
CH3Hg : Methyl mercuric ion
(CH3)2Hg: Dimethyl mercury
CVAAS: Cold vapour atomic absorption spectrometry
EC50: Median effective concentration
GC: Gas chromatography
GSH: Glutathione
Hg0: Elemental mercury
Hg2+: Mercuric ion
HPLC: High-performance liquid chromatography
MerA: Mercuric reductase
MerB: Organomercurial lyase
PCs: Phytochelatins
TT95b: Empirical toxicity threshold
UTmax: Maximum unit toxicity
6
Introduction
Hg - a global environmental pollutant
Sources of Hg pollutants
Mercury (Hg) is a global environmental pollutant that is present in soil, water, air and
biota. Hg enters the environment as a result of natural and human. The naturally occurring
Hg can be released into the atmosphere and then exchanged between the soil and water
systems by the following processes (Ebinghaus et al., 1999):
1. Wind erosion and degassing from Hg mineralized soil and rock formation
2. Volcanic eruptions and other geothermal activities
3. Evasion of Hg from the Earth’s subsurface crust
whereas, anthropogenic sources of Hg can be attributed as follows (Porcella et al., 1996):
1. Combustion of fossil fuels, wood, wastes, sewage sludge and crematories.
2. High temperature processes, e.g. smelting, cement and lime production
3. Manufacturing/commercial activities: e.g. metal processing, gold extraction, Hg
mining, chlor-alkali plants, chemical and instrument industry (Hg chemicals,
paints, batteries, thermometers, process reactants and catalysts).
4. Other sources, e.g. agriculture (pesticides, fertilizers and manure).
Fig. 1 Mercury-cycling in the environment.
7
Current estimates of anthropogenic Hg emission range from about 50 % to 75% of the
total annual Hg emission to the atmosphere (Ebinghaus et al., 1999; Fitzgerald, 1995). The
atmospheric Hg burden has increased by a factor of three during the last 100 years
(Fitzgerald, 1995). The Hg released from both anthropogenic and natural sources is further
distributed in the environment (Fig. 1). The main pathway of Hg transport in the
environment is air-surface exchange with soils, ocean, fresh water and vegetation.
However, other transports like soil-vegetation exchange and water-vegetation exchange
are very important to human beings. The Hg accumulated in vegetation may enter the
human diet either directly or through fish, birds and livestock (Fig. 1). Moreover, the soilvegetation exchange of Hg (Fig. 2) gives a possibility to remove Hg from contaminated
soil by plant uptake.
Fig. 2 The role of terrestrial plants in the biogeochemical cycling of Hg.
Hg speciation in air, water, and soil
The most common gaseous forms of Hg are elemental Hg (Hg0) and dimethyl-Hg
((CH3)2Hg). On a global scale, the atmospheric Hg cycle is dominated by elemental Hg
(generally > 95% of total airborne Hg), whereas only minor amount of other species
(mainly particulate-phase Hg (Hg(p)) have been detected (Stratton and Lindberg, 1995).
Both methyl-Hg and dimethyl-Hg have been detected in ambient air (Bloom and
Fitzgerald, 1988). However, the concentrations are far below those of the inorganic
species. The total Hg concentration in air at background levels is generally 1–4 ng m-3 (Table
1). The atmospheric Hg concentrations in 1990 were 2.25±0.41 and 1.50±0.30 ng m-3,
respectively, in the northern and southern hemispheres over the Atlantic Ocean (Slemr and
8
Langer, 1992), and it was reported as 1.5 ng m-3 at the west coast of Sweden in 2003
(Munthe et al., 2003). The atmospheric Hg concentration is generally higher in urban and
industrial areas, and it was reported to be 600 and 1500 ng m-3 near Hg mines and
refineries (WHO, 2000).
Table 1 Background Hg concentrations in different media and general Hg speciation
Media
Hg concentrations
-3
Hg speciation
0
Air
1–4 ng m
Hg , minor amount of (CH3)2Hg,
CH3Hg-X, and particulate Hg
Ocean
0.3—4.4 ng L-1
Hg-X2, CH3Hg-X,
particulate Hg
Soil
0.003–4.6 µg g-1
Hg2+ and CH3Hg+ (mainly bound to
organic matter, mineral substances,
and sulphide), minor amount of Hg0
Hg0,
and
References
Bloom and Fitzgerald, 1988;
Slemr and Langer, 1992;
Munthe et al., 2003
Bloom and Crecelius, 1983;
Laurier et al., 2004
Steinnes, 1997;
Schuster, 1991
Note: X means OH- or Cl-.
Water contains Hg mainly in the form of Hg2+ as a complex salt bound to dissolved
particles (Table 1). Hg concentrations in rivers, lakes, rain and snow may vary widely
depending on environmental conditions. Mierle (1988) found about 0.3 – 2.2 ng Hg L-1 in
lakes and rivers, and 5 – 40 ng Hg L-1 in brown streams in Ontario, Canada. According to
Bloom and Watras (1988), snow may contain 4 ng kg-1 total Hg and 0.05 ng kg-1 methylHg, and rain samples contain 2–5 ng L-1 of total Hg and 0.15 ng L-1 of methyl-Hg. Ocean
water was considered to have a Hg concentration of 0.3 to 4.4 ng L-1 (Bloom and
Crecelius, 1983; Laurier et al., 2004). In sediments Hg is mainly bound to sulphur as well
as organic matter and inorganic particles (Ullrich et al., 2001).
Mercury levels in surface soils were reported to range from 0.003 – 4.6 µg g-1 on a
global scale (Steinnes, 1997). However, the Hg levels may be rather high in contaminated
sites, e.g. up to 557 mg Hg kg-1 was found in the vicinity of a chlor-alkali plant at Ganjam,
India (Lenka et al., 1992). The chemical state of Hg in soil is apparently related to soil
properties, as well as the chemical character of the water phase, pH, the redox potential
and the presence of organic matter and inorganic agents (Lifvergren, 2001). Hg has
generally a high affinity for organic matter in soil matrix (Schuster, 1991). Hg may form a
stable complex (i.e. HgS) with sulphide in soil, which probably appears incorporated in the
other metal sulphides or in organic matter rather than as crystalline HgS (Benoit et al.,
1999). Hg is also adsorbed to minerogenic substances, e.g. clay minerals and hydrous
oxides of Fe, Al, and Mn. Since Hg is strongly bound to soil constituents, normally, only
trace content of Hg are found in the soil solution (Schuster, 1991). Dissolved forms of Hg
in soil solution are free Hg ions and soluble Hg complex, which are easily utilized by
living organisms. Elemental Hg (Hg0), as well as neutral organic Hg like (CH3)2Hg has
significant vapour pressure. Thus, they are volatile, and a vaporization of Hg can occur
from polluted soil through these Hg species that are only weakly adsorbed on surface of
minerals or organic matter (Lifvergren, 2001). Hg0 in soil can be transformed from Hg2+
9
via abiotic or bacterial reduction. Humic and fulvic acids in soil are able to reduce Hg2+ to
Hg0, and this abiotic reduction process is promoted by sunlight (Allard and Arsenie, 1991;
Xiao et al., 1995). Some bacteria are capable of enzymatically reducing Hg2+ to Hg0 via
mercuric reductase i.e. MerA (Fox and Walsh, 1982; Schiering et al., 1991; WagnerDobler et al., 2000) (Fig. 3-b). Another enzyme, organomercurial lyase (MerB) existing in
some bacteria, catalyzes the cleavage of the carbon-Hg bond of several forms of organic
Hg (Begley et al., 1986) (Fig. 3-a).
Fig. 3 Biochemistry of bacterial Hg detoxification.
a) Organomercurial lyase (MerB) detoxifies organic Hg
(RHg) by catalyzing the cleavage of the carbon-Hg bond;
b) Mercuric reductase (MerA) reduce Hg2+ to Hg0.
In respect to environmental exposures, methyl-Hg compounds present the most critical
concern. Bacteria such as the sulphate-reducing strains are able to methylate Hg2+ to
CH3Hg+ (Compeau and Bartha, 1985; Olson and Cooper, 1976). In general, around 1% of
the total Hg in sediment is converted to methyl-Hg mainly via bacterial activities (von
Burg and Greenwood, 1991). Matilainen et al. (2001) reported that Hg methylation was
most intensive in organic surface layer, especially the living moss in the uppermost 0–16
cm of the soil profiles.
Health risks of Hg
Mercury and its compounds are persistent, bioaccumulative and toxic, and they pose a risk
to both humans and ecosystem. Exposures to Hg, e.g. breathing Hg-contaminated air,
eating Hg-contaminated food products (especially fish), eating and touching Hgcontaminated soil may result in devastating neurological damage, kidney damage, and
even death (Tchounwou et al., 2003; WHO, 1976). Historic and recent industrial activities,
including the mining of gold, silver and Hg itself, have caused Hg contamination of
terrestrial and aquatic ecosystems (Porcella et al., 1996). Hg-contaminated soil is believed
to contribute to human health risks and phytotoxicity of plants. Hg in contaminated soil
may also enter aquatic ecosystems via leaking and cause Hg contamination of fish and
animals that eat fish (Fig. 1). Therefore, the numerous Hg-contaminated sites that exist in
the world have given rise to a great concern for remediation.
Overview of Phytoremediation
A few remediation techniques have been used in practice so far for removal of Hg from
contaminated soil, e.g. washing soil with halogenated substances and heating soil to more
than 600°C (Hempel and Thoeming, 1999). However, these techniques are relatively
expensive and cause further disturbance to the already damaged environment.
Phytoremediation, i.e. using plants to remove pollutants from the environment or to render
10
them harmless, is considered as a promising, cost-effective, and environment-friendly
technology to clean up the contaminated environment (Cunningham and Berti, 1993;
Lasat, 2002; Raskin et al., 1994; Salt et al., 1998). This has led to growing interest in
phytoremediation from governments, organizations, and industries. The world
phytoremediation market was estimated at $55–$103 million dollars in 2000, reaching
$214–$370 million in 2005 (D. Glass Associates Inc., 1999).
Advantages and disadvantages of phytoremediation
Macek et al. (2000) gave a comprehensive review of the advantages and disadvantages of
phytoremediation. The main advantages of phytoremediation are:
•
•
•
•
•
•
Low operating costs
Far less disruptive to the environment
In situ application avoids excavation.
Large-scale clean-up operations
A relatively easy process with available equipment and supplies generally used in
agriculture
High probability of public acceptance
Like any other method of environmental remediation, phytoremediation has its
disadvantages:
•
•
•
•
•
Slower than some other alternatives to restore an area
Limit of the climatic and geological conditions of the contaminated site, e.g.
temperature, altitude, soil type, and accessibility to agricultural equipment
Biological methods are not capable of 100% reduction of contaminants
Formation of vegetation may be limited by extremes of environmental toxicity
Need to take care of the accumulators after remediation to avoid reemission
Processes and technologies of phytoremediation of heavy metals
There are a number of different types of phytoremediation processes, which can be applied
to both organic and inorganic pollutants present in soil, water and air (Cunningham and
Ow, 1996; Cunningham et al., 1995; Raskin et al., 1997; Salt et al., 1995, 1998). Five of
them are relevant to the phytoremediation of Hg, which is one of the most difficult heavy
metals to be removed by means of phytoremediation. These five subsets of
phytoremediation are termed as phytoextraction, phytovolatilization, phytostabilization,
rhizofiltration, and foliage filtration.
•
Phytoextraction is the use of pollutant-accumulating plants to remove metals from
soil by concentrating them in the harvestable parts (Salt et al., 1995). In phytoextraction, metal-tolerant plants with high metal accumulation, high translocation of metal
into the shoots and high biomass production are used. To avoid contamination of air,
11
plants used should not release large amounts of Hg into the atmosphere. Interest in
phytoextraction has grown significantly following the identification of metal
hyperaccumulator plants (Lasat, 2002). Hyperaccumulators are the species capable
of accumulating metals at levels 100-fold greater than those typically measured in
shoots of common nonaccumulator plants (McGrath and Zhao, 2003). More than
400 plant species have been identified as natural metal hyperaccumulators, e.g.
Thlaspi spp. could accumulate up to 31000 µg g-1DW of Ni and 43710 µg g-1DW of
Zn, however, no Hg hyperaccumulators were found (Reeves and Baker, 2000).
•
Phytovolatilization involves the use of plants to take up pollutants from soil,
transforming them into volatile forms and transpiring them into the atmosphere.
Transgenic plants expressing genes merA and merB could convert hazardous
methyl-Hg and ionic Hg to the less toxic volatile elemental Hg, which is released to
the air (Bizily et al., 1999, 2000; Rugh et al., 1996, 1998). However, the Hg released
into the atmosphere is likely to be recycled and deposited back into lakes and
oceans, repeating the production of methyl-Hg via bacterial methylation. Therefore,
phytovolatilization of Hg is not recommended.
•
Phytostabilization is the use of plant roots to reduce the mobility or bioavailability of
pollutants in the environment (Pulford and Watson, 2003). The plants used for
phytostabilization should have efficient root-accumulation of metals, low
translocation of metals to the shoots, and large root system. The potential use of
trees, especially willow, for phytostabilization of heavy metal-contaminated land has
received increasing attention over the last 10 years (Pulford and Watson, 2003).
•
Rhizofiltration is the use of plant roots to absorb, concentrate, and precipitate heavy
metals from water and aqueous waste streams (Ensley, 2000).
•
Foliage filtration is the use of plants to remove pollutants from air by uptake via the
leaves.
All the phytoremediation processes mentioned above require that the plants used are
able to tolerate Hg. Moreover, the phytoremediation process that can be applied in practice
is determined by the ability of plants in accumulating, translocating, and volatilizing Hg.
Plant interaction with Hg
Plants are capable of extracting a variety of metal ions from their growth substrates,
including Hg. Many studies have showed that plant roots accumulate Hg when they were
exposed to Hg-contaminated soils (Bersenyi et al., 1999; Coquery and Welbourn, 1994;
Lenka et al., 1992; Kalac and Svoboda, 2000; Ribeyre and Boudou, 1994). Laboratory
studies showed that plant roots absorbed Hg from solution and roots accumulated much
greater amount of Hg than shoots (Beauford et al., 1977; Cavallini et al., 1999; Godbold
and Hütterman, 1988). Both field and laboratory studies have demonstrated that plants
accumulate more Hg when it is introduced in organic form than in inorganic form
(Godbold, 1991; Godbold and Hütterman, 1988; Ribeyre and Boudou, 1994).
12
Leaves can absorb gaseous Hg via stomata, which has been shown in previous
laboratory studies (Browne and Fang, 1978; Cavallini et al., 1999; Du and Fang, 1982,
1983). Du and Fang (1982) reported that uptake of Hg0 by the leaf increased with
increasing Hg vapour concentration, temperature, and illumination. Leaves can also absorb
Hg after deposition of particulate Hg on the leaf surface (de Temmerman et al., 1986;
Fernández et al., 2000) and release gaseous Hg into the atmosphere (Siegel et al., 1974;
Kozuchowski and Johnson, 1978). Furthermore, Hanson et al. (1995) reported that at low
external Hg concentrations in the air, the release of Hg from leaf to air was higher than the
leaf Hg absorption from the air in the tree species Picea abies L. Liriodendron tulipifera
L., Quercus alba L., and Acer rubrum L.. Similar results were also found by Ericksen and
Gustin (2004). This evidence suggests that foliage can manage both uptake and
volatilization of gaseous Hg.
All physiological and biochemical processes in plants may be negatively affected by
Hg when plants are exposed to Hg-contaminated soil, water or air (Patra and Sharma,
2000). Elemental Hg (Hg0) does not react with most biomolecules unless first oxidized to
Hg2+, and this may be catalytically driven by peroxidase or catalase (Du and Fang, 1983;
Ogata and Aikoh, 1984). Hg cations have a high affinity for sulphydryl (-SH). Because
almost all proteins contain sulphydryl groups or disulphide bridges (-S-S-), Hg can disturb
almost any function in which proteins are involved in plants (Clarkson, 1972). Organic Hg
is 1–2 orders of magnitude more toxic to some eukaryotes and is more likely to
biomagnify across trophic levels than ionic Hg (Hg2+) (Liu et al., 1992; Bizily et al., 2000).
The biophysical behaviour of organic Hg is thought to be due to its hydrophobicity and
efficient membrane permeability (Braeckman et al., 1998). Hg compounds can also bind to
RNA, several synthetic polyribosomes, and DNA (Katz and Santilli, 1962; Kawade, 1963;
Cavallini et al., 1999). Hg is known to affect photosynthesis, mineral nutrient uptake, and
transpiration (Barber et al., 1973; Godbold, 1991, 1994; Godbold and Hütterman, 1988;
Patra and Sharma, 2000). Plants can generally sequester toxic ions in complexes at the
cytoplasm to defend against their phytotoxicity. Glutathione (GSH)-related phytochelatins
(PCs) with the general structure (γ Glu-Cys)nGly (n=2-11) are the most dominant
molecules found so far to sequester the metal ions in cytoplasm and then transport them to
vacuoles (Rajesh et al., 1996; Zenk, 1996). Rajesh et al. (1996) reported that the strength
of Hg(II) binding to glutathione and phytochelatins ranked in order as follows: γ Glu-CysGly < (γ Glu-Cys)2Gly < (γ Glu-Cys)3Gly < (γ Glu-Cys)4Gly. Suhadra et al. (1993) found
that, compared with normal seedlings, those from Hg-treated seeds exhibited a larger
amount of nonprotein SH, indicating the possible involvement of phytochelatins in the Hgreduced adaptive response. Organic acids (e.g. citrate) and amino acids (e.g. histidine)
existing in cytoplasm may also complex metal ions and reduce their toxicities to plants
(McGrath and Zhao, 2003).
Although quite a few reports, as mentioned above, have involved Hg toxicity,
tolerance, uptake and release of plants, systematic study in view of the potential
application of plants in phytoremediation is far from sufficient.
13
Aim of the present study
The aim of this thesis was to study the fate of Hg in terrestrial plants and the effect of
plants on Hg geochemistry in contaminated soil and, thus to clarify the potential use of
plants for phytoremediation of Hg. This aim was achieved by fulfilling the following
objectives:
14
¾
To compare the ability of various plant species to take up and release Hg to the
atmosphere via the leaves. The release of Hg into air by plants is a concern when
using phytoremediation in practice (Paper I).
¾
To determine whether willow is tolerant to Hg, and if it accumulates and
translocates large quantities of Hg to the shoots. It was hypothesised that willow
clones with high tolerance, accumulation, and translocation of Hg to the shoots
could be selected for future phytoextraction experiments (Paper II).
¾
To compare the accumulation and distribution of Hg in willow after exposure to
methyl-Hg and inorganic Hg. Moreover, to investigate the possibility of
methylation inside the willow plant. The methyl-Hg present in the environment is
a most critical concern regarding human health.
¾
To investigate i) if a Hg-tolerant willow clone is able to accumulate and
translocate large quantities of Hg to the shoots from a Hg-contaminated soil, ii) if
the plant accumulation correlate to the Hg extracted by various chemical
extractants, and iii) if willow is able to modify the chemical forms of Hg in the
soil (Paper III).
¾
To find out whether iodide can increase the bioavailability of Hg in soil and,
therefore enhance the phytoextraction of Hg-contaminated soil with willows by
investigating i) sensitivity of willow to iodide, ii) whether the plant can efficiently
accumulate Hg along with iodide in hydroponics, iii) whether iodide addition
increase the accumulation of Hg from contaminated soil, and iv) whether iodide
addition increases translocation of Hg to the shoots (Paper IV). The proposed low
bioavailability of Hg in contaminated soil is a restricting factor in phytoextraction
of Hg.
Comments on materials and methods
Plant materials
Willow (Salix spp.) was chosen to study the Hg fate in plants as well as the
phytoremediation of Hg-contaminated soil in this thesis. As various clones of Salix
viminalis showed different capacities in accumulation, tolerance and translocation of Zn,
Cd and Cu (Greger and Landberg, 1999), it was hypothesised that this might also be the
case for Hg. Moreover, willow has a large root system, which is capable of high hydraulic
pumping pressures. Some willow clones have high biomass production, which can
produce 70 tons of stem per 3-year rotation under favourable conditions (Labrecque and
Teodorescu, 2003), and they are currently being used for bioenergy production in Sweden.
It was hypothesised that high-biomass-producing willow clones with high tolerance,
accumulation, and translocation of Hg to the shoots could be selected for phytoextraction,
whereas those with high tolerance and root-accumulation of Hg, and low translocation of
Hg to the shoots could be used for phytostabilization. Moreover, those willow clones with
low translocation of Hg to the shoots could be recommended to use for bioenergy
production, which in turn would reduce the Hg emission during combustion of the shoots.
The other plant species, i.e., garden pea (Pisum sativum L. cv Faenomen), spring wheat
(Triticum aestivum L. cv Dragon), sugar beet (Beta vulgaris L. cv Monohill), oil-seed rape
(Brassica napus L. cv Paroll) and white clover (Trifolium repens L.) were used to compare
the ability of various plant species in taking up and releasing Hg to the atmosphere via the
leaves. They are all crops growing as monocultures in large fields in Sweden.
Consequently, if they release Hg their contribution to the atmospheric Hg pool may be
significant.
Plant chamber systems and Hg traps
A set of plant-chamber systems was established to study the Hg accumulation and the
translocation of Hg to the shoots in hydroponics. Hg can easily volatilize from solution
into air (Paper IV) and leaves can absorb gaseous Hg (Du and Fang, 1982; Wang and
Greger, unpublished). A transpiration chamber system (Fig. 4) was constructed to evaluate
the Hg accumulation, translocation and volatilization via leaves, which efficiently
prevented leakage of air into the upper cylinder from the chamber below (Paper I).
Another plant-chamber system with a gaseous Hg generator (i.e. a tube system with a drop
of metallic Hg inside) was used to study uptake of Hg via the shoot (Fig. 5; Wang and
Greger, unpublished).
Hopcalite traps (granules of Cu and Mn oxides) and gold traps were used to analyse Hg
concentration in air (Paper I; Wang and Greger, unpublished). The Hopcalite trap, with a
high concentration working range (µg level), can analyse high Hg-contaminated air,
whereas the gold trap, with a much lower detection limit (pg level), was used to analyse
low Hg-contaminated air.
15
Fig. 4 The transpiration chamber (design II) used to study the volatilization of Hg from shoots
(Paper I).
Fig. 5 The plant-chamber system used to study uptake of Hg from air via shoots (Wang and
Greger, unpublished).
16
Weibull function
To evaluate the differences in Hg tolerance among willow clones, a modified Weibull
model (Taylor et al., 1991) was used in Paper II to compare the dose-response curves. The
modified Weibull function has been shown to be an excellent tool to compare doseresponse curves and estimate some important parameters such as the empirical toxicity
threshold (TT95b), the maximum unit toxicity (UTmax) and EC50. TT95b is the concentration
of Hg where growth is reduced by 5%, EC50 is the concentration of Hg where growth was
reduced by 50% and UTmax indicates the value of the maximum slope of the dose-response
curves (Fig. 6). Generally, a plant with higher values of TT95b and EC50 and a low value of
UTmax means it is less sensitive to Hg than the plants with lower values of TT95b and EC50
and a higher value of UTmax.
Fig. 6 Illustration of the empirical toxicity threshold (TT95b), the maximum unit toxicity
(UTmax) and EC50 in a dose-response curve. The growth of the shoot decreases with increased
Hg concentration in solution when roots were exposed to various concentrations of HgCl2.
* dy/dx indicates the slope of the curve.
Hg analyses
All plant samples that were analyzed for total Hg content were wet-digested in
HNO3:HClO4 (7:3,v/v) for 18 hours in a heating programme with the highest temperature
set at 180°C (commonly 225°C for other metals). The lower temperature was used to
avoid the volatilization of Hg during digestion.
The plant samples used for analyzing organomercury compounds in the Hg
methylation study (Wang and Greger, unpublished; Wang et al., 2002) were extracted with
a method by Ortiz et al. (2002). Samples were subsequently treated with different reagents
17
such as 1.5M KBr/1M CuSO4 in H2SO4 medium, 0.01 M NaS2O3 and 0.5M CuCl2 and
finally, the organic species were extracted in dichloromethane.
To evaluate the changes of various Hg soil fractions in soil induced by willow, a
sequential soil extraction process according to Martin-Doimeadios et al. (2000) was used
in Paper III. The sequential soil extraction method can provide a view of Hg associated
with different soil components and gives much broader information on Hg states in soil
than single soil extractions.
The total Hg content in the solutions, extracts, and digests were analyzed by cold
vapour atomic absorption spectrometry (CVAAS), except that in xylem sap, whereas
methyl-Hg content was analysed by gas chromatography-pyrolyser-atomic fluorescence
spectrometry (GC-pyrolyser-AFS).
Xylem sap was obtained with a capillary glass tube from the shoot xylem with a high
pressure of N2 (5 bar) that being given through the initial cutting. Because a low volume
of xylem sap (1–5 µl) was achieved and its Hg concentration was not high enough,
CVAAS could not be used to analyze the Hg concentration in the xylem sap. Therefore,
electrothermal AAS in combination with a gold-trap was used to analyze Hg
concentrations in the xylem sap. The measurements were carried out with a selfconstructed apparatus consisting of a sample vessel (quartz glass tube), a gold trap, and a
quartz cell for measurement with AAS (Fig. 7). The liquid Hg sample was injected into the
sample tube. While heating the sample tube over 800 ºC the Hg-compounds were
converted to Hg0 by thermal reduction. Thereafter, Hg0 carried by N2 flow was trapped in
the gold trap. Hg collected in the gold trap was then thermally desorbed (800 ºC) and
analyzed with AAS. This technique was also used to analyze the Hg content in a small part
of the root (Wang and Greger, unpublished).
Fig. 7 Scheme of the Hg-pyrolysis apparatus (Wang and Greger, unpublished).
18
Results and discussion
Hg accumulation and distribution in six terrestrial plant species
Our studies show that the six investigated plant species (i.e. willow, garden pea, spring
wheat, sugar beet, oil-seed rape and white clover) were able to accumulate Hg in
hydroponics. About 208 to 407 µg (g DW)-1 of Hg were found in the roots after cultivation
in 1 µM HgCl2 for 2 or 3 days (Paper I). Only a slight but not significant difference in Hg
accumulation was observed between species. The lowest Hg concentration was found in
roots of wheat (208 µg g DW-1), whereas the highest concentration was found in oil-seed
rape (407 µg gDW-1) after the roots had been cultivated in 1 µM HgCl2 for 2 days. A
negative correlation between root dry weight and Hg concentration was found among the
investigated species (Paper I), which indicates that species differences in total root
biomass may have affected Hg accumulation, i.e. a high biomass may cause lower Hg
concentration due to dilution. Similar relationships have also been found for Cd
concentration and biomass production in Scots pine in a previous study (Ekvall and
Greger, 2003).
Furthermore, all investigated plant species translocated Hg from roots to shoots when
the roots were exposed to Hg. However, the translocation rates were low, i.e. only 0.2–
2.5% of the Hg accumulated by the roots was translocated to the shoots in the six
investigated plant species, among which, pea had the highest translocation of Hg in the
shoots and clover had the lowest. This result is consistent with previous studies on Pisum
sativum L., Mentha spicata L., and Picea abies (L.) Karst (Beauford et al., 1977; Godbold
and Hütterman, 1988). The low translocation of Hg to the shoots is probably due to the
high affinity of roots for Hg. The Hg trapped in the roots was mainly bound to root cell
walls, about 80% of accumulated Hg was bound in the cell wall of the willow roots (Paper
II).
Hg accumulation and distribution in eight willow clones
There was no difference detected in Hg accumulation and translocation of Hg to the shoots
among the eight willow clones after being cultivated in 0.5 µM HgCl2 for 3 days (Paper
II). This result was not in line with previous studies on some other heavy metals by Greger
and Landberg (1999) and Greger et al. (2001), who have reported a distinct variation in
accumulation and translocation of Zn, Cd and Cu among various willow clones. Willow
roots efficiently accumulate Hg in hydroponics, the Hg concentrations of roots ranged
from 216 to 274 µg (g DW)-1, and only 0.45 to 0.62% of the total Hg accumulated via
roots was translocated to the shoots in the eight investigated willow clones after 3 days of
cultivation in 0.5 µM HgCl2 (Paper II). These translocation rates were lower compared
with those of some other heavy metals (Greger and Landberg, 1999).
19
Accumulation and distribution of various Hg species in willow
Willow clone Björn was used in this study. The short-term Hg-accumulation study (Paper
IV; Wang and Greger, unpublished) showed that Hg accumulation in the roots of willow
decreased according to its species in following order: CH3HgCl > HgCl2 ≈ Hg(NO3)2 >
HgI2 (Fig. 8). Differences were found in the effective Hg accumulation between CH3HgCl
and the other three Hg species after 3 days cultivation (Table 2). Plants could accumulate
more methyl-Hg than inorganic Hg, which has also been shown in previous studies
(Godbold, 1991; Godbold and Hütterman, 1988; Ribeyre and Boudou, 1994). Because Hg
mainly forms stable complex, i.e. HgI42- with a stability constant of 29.8 in HgI2 solution
(Wasay et al., 1995), willow roots accumulated Hg from HgI2 solution slower than from
Hg(NO3)2 solution (Fig. 8; Paper IV). However, the effective Hg accumulation from HgI2
solution (i.e. 66.9%) was similar to that from Hg(NO3)2 (i.e. 65.0%) after 3 days of
cultivation (Paper IV). Because both the root accumulation and the bacterial volatilization
of Hg were faster in Hg(NO3)2 than in HgI2 solution (Fig. 8; Paper IV), the decrease of Hg
concentration was faster in Hg(NO3)2 solution than that in HgI2 solution. After 3 days of
cultivation, there was only 5% of the total Hg left in the Hg(NO3)2 solution, whereas 26%
was left in the HgI2 solution (Paper IV). As shown above, both the faster reduction of Hg
accumulation rate and the faster loss of Hg by the bacterial volatilization in Hg(NO3)2
solution may explain why the effective Hg accumulation from Hg(NO3)2 solution was
similar to that from HgI2 solution (Paper IV).
Table 2 Accumulation and distribution of Hg in willow after being cultivated in 800 ml
nutrient solution with 1 µM CH3HgCl, HgCl2, Hg(NO3)2, and HgI2, respectively, for 3 days.
n=3, ± SE (data from Paper IV and from Wang and Greger (unpublished))
Effective Hg Distribution Hg in Leaves Hg in branches
Treatments accumulation to the shoot µg (g DW)-1 µg (g DW)-1
%
%
Hg in roots
µg (g DW)-1
Hg in xylem
sap †
µM
0.23±0.07 a
—
—
—
0.05±0.01 a
0.05±0.01 a
CH3HgCl
79.2±7.7 a
0.47±0.05 a
0.58±0.04 b
1.03±0.05 c * 966.20±76.71 c 0.12±0.04 a
HgCl2
62.1±2.4 b
0.49±0.02 a 0.68±0.04 b *
0.48±0.02 b
515.66±35.50 b 0.10±0.04 a
Hg(NO3)2
65.0±2.8 b
0.48±0.03 a 0.76±0.09 b *
0.53±0.04 b
464.01±38.62 b
—
HgI2
66.9±3.7 b
0.46±0.02 a
0.56±0.10 b
398.12±42.80 b
—
No Hg
0.61±0.07 b
a,b,c Different letters in each column denote significant difference at p ≤ 0.05.
* Significant difference between leaves and branches at p ≤ 0.05.
† Xylem sap was collected by using capillary glass tube after plants being treated with Hg for 12 h.
20
Fig. 8 Accumulation of Hg in willow roots during the cultivation of plants in 1 µM CH3HgCl,
HgCl2, Hg(NO3)2, and HgI2, respectively, for 4 h. The Hg solution was changed every 30 min.
The Hg accumulation rate is shown as the slope of the line between the two harvest times. n=3,
± SE (data from Paper IV and from Wang and Greger (unpublished)).
Willow could efficiently accumulate Hg from the solutions of various Hg species, i.e.
more than 60% of Hg added in all solutions was accumulated by the plants during 3 days
(Table 2). However, no differences in translocation of Hg to the shoots were found among
the investigated Hg species (Table 2). Nonetheless, the distribution patterns of Hg in the
shoots were found to be variable among the four investigated Hg species. Hg
concentration in the branches was higher than that in the leaves of plants that had been
cultivated in CH3HgCl solution for 3 days, whereas it was the opposite in the cases of
21
HgCl2 or Hg(NO3)2. An explanation for these differences is that methyl-Hg might cross
the plasma membrane easier and therefore it can be loaded more to the surrounding cells
of xylem vessels and the wall of vessels (probably especially to the lignin) during the
translocation from roots to leaves. Methyl-Hg has been shown in previous studies to more
easily cross the plasma membrane than inorganic Hg (Braeckman et al., 1998).
According to our studies and the literature, all previously investigated plants have low
translocation of Hg to the shoots (Table 2; Papers I–IV; Beauford et al., 1977; Godbold
and Hütterman, 1988). It seems that the majority of the total accumulated Hg is trapped in
the roots and that only a minor amount can be translocated to the shoots. Ion uptake
mainly occurred at the root tip, prior to the formation of the Casparian band (Fig. 9),
which is a zone allowing apoplasmic transport of heavy metals into the stele (Marschner,
1995). The Casparian band, i.e. hydrophobic incrustations (suberin) in the radial and
transverse walls of the endodermis, constitutes an effective barrier against passive
movement of heavy metals into the stele, thus influencing metal accumulation (Fig. 9-c;
Lux et al., 2004; MacFarlane and Burchett, 2000). When the whole wall of the endodermis
cell is deposited with suberin lamellae, only a symplastic transport of ions from cortex
cells enables ions to enter the stele (Fig. 9-b). The distance between formation of suberin
lamellae and the root tip influences the translocation of metals, e.g. Cd, to the shoots (Lux
et al., 2004). Our studies showed that accumulation of Hg from either HgCl2 or CH3HgCl
solution was higher at the root tip of willow prior to formation of the Casparian band than
after formation, when a whole root had been cultivated in Hg solution for 4 hours (Fig.
10). However, only trace concentrations of Hg were found in the part of the root with the
Casparian band, i.e. less than 1% of the Hg concentration in the root tip, when only the
root tip was treated with either HgCl2 or CH3HgCl solution (Fig. 10). This indicated that
only a minor amount of the Hg accumulated in the root tip was translocated to the rest of
the root. It seems that most of the Hg accumulated from the Hg solutions was trapped in
epidermis and cortex, and mainly absorbed by their cell walls when passing across the
cortex to the stele. Consequently, the Hg concentration found in xylem saps was low, i.e.
0.12µM and 0.10µM, respectively, in terms of CH3HgCl and HgCl2 (Table 2; Wang and
Greger, unpublished), which was much less than the original Hg concentration in the given
solution (i.e. 1 µM).
22
Endodermis with
Casparian band
Stele
Cortex
d
Fig. 9 Illustration of the Casparian band in root tip of willow. a) Diagram of longitudinal section of
a root tip (adapted from Esau (1953)); b) The diagram represents only a symplastic transport of ions
from cortex cells to stele when the endodermis cell wall is deposited with suberin lamellae (adapted
from Mauseth (1988)); c) The diagram represents that the Casparian band is a barrier of apoplastic
movements of water and solutes from cortex to stele (adapted from Mauseth (1988)); d)
Fluorescence microscopic picture of a hand section of willow root stained with berberine,
bar = 50 µm (Wang and Greger, unpublished).
23
Fig. 10 Distribution of Hg in willow roots after either the whole root or the root tip had been
treated with 1 µM CH3HgCl and HgCl2, respectively, for 4 h. Willow roots used were 3-weeks
old with the Casparian band starting at 3–10 mm above root tip. Willow nutrient solutions (NS)
with or without Hg were dropped onto the roots at a flow rate of 3 ml min-1. * Relative Hg
concentrations were calculated as Hg concentration (pg per 1.5-mm long root cutting) in
different cutting positions in relation to that in cutting position 1. Background concentrations of
Hg in roots cultivated without Hg were subtracted from concentrations in treated roots before
the calculation. n=3, ± SE (Wang and Greger, unpublished).
Screening of Hg accumulators in a Hg-contaminated site
In order to find Hg accumulators, fifteen plant species were collected from a chlor-alkali
plant area at Bohus site in the vicinity of Gothenburg (Sweden) with Hg concentrations of
3.6–50.0 µg g-1DW in soil. Hg concentration in shoots ranged from 0.09 to 1.84 µg g-1DW
(Table 3). Western thistle (Cirsium arvense) was found to have the highest Hg
concentration in both shoots and roots among 15 investigated plant species. This plant
specie was then used to compare the Hg accumulation with willow in a field trial. As a
result, the Hg concentration in the shoots of western thistle was found to be higher than
that of willow, whereas Hg concentration in the roots was lower (Table 4). However, these
differences were not large.
24
Table 3 Screening of plant species in Hg accumulation from a Hg-contaminated site in the
vicinity of a chlor-alkali plant in June, 2002
Hg in plants, µg (g DW)-1
Shoots
Roots
Hg in soil
µg (g DW)-1
Cirsium arvense
1.84
6.4
30.3 a
Populus tremula
0.88
1.98
22.5 a
Plantaginaceae spp.
0.86
—
50.0 b
Plantago major
0.78
1.78
14.9 a
Calamagrostis arrundinacea
0.75
2.23
50.0 b
Rumex spp.
0.72
—
16.4 b
Chamomoilla suvaveolens
0.65
—
10.6 b
Trifolium. medium
0.46
—
16.4 b
Potentilla anserina
0.41
—
10.6 b
Epilobium angustifolium
0.34
0.98
16.4 b
Aegopodium podagraria
0.24
—
3.2 b
Salix cinerea
0.26
—
3.2 b
Betulaceae pendula
0.24
—
16.4 b
Vicia cracca
0.22
—
16.4 b
Equisetum arvense
0.09
0.21
3.62 a
Plants
a The Hg concentration in the soil attached to the roots;
b The average of the Hg concentration in the soil of 2–3 samples from the sampling location.
Table 4 Mercury concentration in soil near roots and accumulation of Hg in plants after being
cultivated in aged Hg-contaminated soil in the vicinity of a chlor-alkali plant for one growing
season. The initial Hg concentration in the soil was 23.2 ±0.7 µg (g DW)-1. n=3, ± SE
Plants
Hg in plants, µg (g DW)-1
Shoots
Willow
Western Thistle
Roots
0.72±0.12
1.02±0.09
Hg in soil near roots
µg (g DW)-1
*
4.20±0.44
2.51±0.31
*
23.3±0.9
22.2±2.1
* Significant difference between plant species at p ≤ 0.05.
The Hg concentration in the roots of willow in the field trial was much lower than that
in hydroponics, although the Hg concentration in soil was much higher (Tables 2, 4).
Similar results were found in the pot experiment (Paper III), which was supposed to be due
to the low bioavailability of Hg in soil. However, relatively higher Hg content in the
shoots was found in field trials (Table 4; Paper IV), which was apparently not due to the
25
translocation of Hg from root accumulation but the leaf uptake of Hg from air. As the air
Hg concentration at the site measured was up to around 250 ng m-3 (Wängberg et al.,
2003), the willow leaf could absorb Hg from Hg-contaminated air and contribute to the Hg
content in the shoot (Fig. 11).
Hg exchange between plants and air
Willow leaves were able to absorb gaseous Hg from air. The concentration of Hg in leaves
increased with increasing exposure time to Hg-contaminated air (Fig. 11). Similarly,
Barghigiani and Bauleo (1992) investigated Abies alba grown near a mining area and found
that the Hg content in needles increased with age. The average rate of the Hg uptake in
willow leaves was found to be 0.45 µg Hg (g dw)-1d-1 in a 12-day period when shoots were
exposed to 7.2 µg m-3 Hg-contaminated air (Fig. 11). Moreover, uptake of Hg0 by the leaf
was found to increase with increasing Hg vapour concentration (Wang and Greger,
unpublished). Hg was also detected in the branches, which is thought to be due to the
translocation of Hg from leaf uptake or direct absorption of Hg by the surface of branches
from air. About 8% of the total Hg that passed though the chamber was absorbed by the
shoots. The Hg concentration in the roots was also analyzed, however, no significant
difference was found between Hg treated and control. This is probably due to a negligible
translocation of Hg from shoots to roots.
Fig. 11 The content of Hg in willow leaves and branches when the shoots were exposed to
gaseous Hg in a plant chamber. The airflow (45 L h-1) passed through a tube system with a
drop of metallic Hg into the plant chamber (Fig. 5). The Hg concentration in the passing air
before entering the plant chamber was measured as 7.2 µg m-3. Plant samples were collected
after 1, 3, 6 and 12 days. ± SE, n=3 (Wang and Greger, unpublished).
26
Leaves can absorb gaseous Hg via stomata, which has been shown in previous
laboratory studies (Browne and Fang, 1978; Du and Fang, 1982; Cavallini et al., 1999).
When gaseous Hg (mainly Hg0) enters into cells of leaves it may be catalytically oxidized
to Hg2+ by peroxidase or catalase and, thus bind to biomolecules (Du and Fang, 1983;
Ogata and Aikoh, 1984). On the other hand, Hg0 may also be abiotically oxidized to Hg2+
in the air (Lindberg and Stratton, 1998) and thereafter bind to negative charges in the
surfaces of leaves or branches.
Whether leaves can release gaseous Hg into the atmosphere or not is an important
factor in the use of plants for phytoremediation of Hg, because contamination of the air has
to be taken into account. By using a transpiration-chamber system, no Hg was found to be
released from shoots into the air in any of the investigated plant species, i.e. garden pea,
spring wheat, sugar beet, oil-seed rape, white clover and willow, after their roots had been
cultivated in 1 µM HgCl2 for 2, 3 and 4 days, respectively (Paper I). Moreover, the same
results were obtained when willow was cultivated in solutions with 1 µM of the Hg
species, i.e. CH3HgCl, HgNO3 or HgI2 (Wang and Greger, unpublished). These results
neither corroborate with those of Kozuchowski and Johnson (1978), who found gaseous
emissions of Hg from aquatic plants in nature, nor with those of Ericksen and Gustin
(2004) and Hanson et al. (1995), who showed an exchange of Hg between leaf surfaces of
trees and air. The different results may be due to variations between plant species in the
property to release Hg. However, we suggest that the reason for these differences may be
due to the experimental set-up. It is difficult to control the emission of Hg originating from
bacterial activity in soil and water, because ionic Hg (Hg2+) is easily converted to
elemental Hg (Hg0) by bacterial activity (Paper IV; Fox and Walsh, 1982; Schiering et al.,
1991; Wagner-Dobler et al., 2000). Therefore, the proposed Hg emitted from leaves might
be released without passing the plants. Our two chamber designs showed that a middle part
(Fig. 4) is needed to capture the gaseous Hg evaporated from the solution (Paper I). In the
absence of the middle part, an increased amount of Hg was found in the Hopcalite trap,
whereas in the presence of the middle part, no increased Hg level was found in the
Hopcalite trap (Paper I). In addition, it is also very important to keep higher air pressure in
the upper cylinder to prevent any possible leaking of air into the upper cylinder from the
chamber below.
In order to know whether Hg affected the transpiration stream or not, which in turn
would decrease the translocation and release of Hg, water transpiration was measured.
There was no effect of Hg detected on the transpiration of water or on stomata (Paper I).
Therefore, Hg release could occur if the metal followed the transpiration stream to the
shoots and directly evaporated from stomata. Our data showed that there was no Hg release
from leaves into air, whereas in the case of willow the plant contained 90 µg Hg (Paper I).
This apparently shows that Hg was trapped in the plant tissues and no Hg was released
from stomata. However, such Hg-containing tissues might lose their physiological
functions, die and become debris, and bacterial activities could then convert Hg2+ to Hg0
that will eventually be emitted into air. Therefore, we can assume that a release of Hg may
be found in a prolonged exposure of plants to Hg, however, not via stomata.
27
Sensitivity of willow to Hg
Our test showed that Hg reduced the growth of roots and shoots of willow when it was
cultivated in HgCl2 solutions (Paper II). A large variation in sensitivity to Hg was found
among willow clones. The EC50 value of the six willow clones ranged from 0.28 to 1.55
µM in terms of dry weight of shoot mass and from 0.29 to 1.95 µM in terms of dry weight
of root mass (Table 5; Paper II). Toxicity threshold (TT95b) and maximum unit toxicity
(UTmax) also showed large differences among the clones (Table 5). The variation in
sensitivity among willow clones has been found to occur with some other metals as well
(Landberg and Greger, 1996; Greger and Landberg, 1999; Greger et al., 2001). The values
of Weibull parameters in terms of roots were shown to be different from those in terms of
shoots (Table 5). Similar results were reported by Österås et al. (2000). This difference is
supposed to be due to the differences in the affinity of Hg for key enzymes of
physiological pathways and the tolerant mechanisms between roots and shoots (Tommy
Landberg, pers. comm.). Among the six tested plants, clone 88-31-7 was the most
sensitive according to TT95b, UTmax, and EC50 in both roots and shoots. The clone with
highest tolerance, however, depended on which parameter was used. According to TT95b,
clone Björn had the highest root tolerance while clone 88-11-4 had the highest shoot
tolerance. However, according to UTmax or EC50, 88-11-4 had the highest root tolerance,
and Björn had the highest shoot tolerance.
Table 5 Interpretation of the differences in Hg toxicity among six willow clones using the
modified Weibull frequency distribution to model dose responses to Hg † (n = 3, ±SE) (data
from Paper II)
Clone
Björn
Tora
88-11-4
Orm
78183
88-31-7
Part
R2
shoot
root
shoot
root
shoot
root
shoot
root
shoot
root
shoot
root
0.99
0.84
0.79
0.86
0.97
0.71
0.95
0.88
0.91
0.84
0.94
0.79
TT95b‡
EC50‡
–––––––––µM–––––––––
0.22
1.55
0.61
1.28
0.21
1.44
0.47
1.16
0.40
1.04
0.23
1.95
0.09
0.95
0.21
0.81
0.37
0.93
0.33
1.01
0.05
0.28
0.02
0.29
UTmax‡
%
36.0
95.6
38.9
86.6
89.9
28.4
60.9
85.9
104.6
82.0
212.2
206.5
† Calculations are based on the dry weight of shoot and root mass.
‡ The terms TT95b and EC50 indicate the HgCl2 concentration where growth was reduced by 5 and 50%,
respectively, whereas UTmax is the maximum unit toxicity (% of growth response/µM).
Our study showed that the transpiration of water in the plants decreased by 49% after
willow had been treated with 1 µM CH3HgCl for 3 days, whereas it was not influenced in
28
the case of HgCl2 (Wang and Greger, unpublished). Previous studies have shown that
methyl-Hg is more toxic to plants (Godbold, 1991, 1994; Godbold and Hütterman, 1988),
which is thought to be because methyl-Hg was easier to pass the plasma membrane
(Braeckman et al., 1998). Methyl-Hg is also much more toxic to human beings and
animals (Liu et al., 1992), thus, the possible methylation present in plants might be a
potential risk to the ecosystem when using plants for phytoremediation. However, our
studies showed that no methyl-Hg was found in leaves, branches, or roots, after willow
had been cultivated in 1 µM HgCl2 for 3 days under the detection limit of 2.5 µg kg-1 DW
(Wang and Greger, unpublished).
Toxic effects of gaseous Hg on leaves were also observed in Hg-contaminated air
(Wang and Greger, unpublished), which was supposed to be due to Hg0 in the air being
oxidized into Hg2+ biotically or abiotically and thus, reacting with biomolecules (Du and
Fang, 1983; Ogata and Aikoh, 1984; Lindberg and Stratton, 1998). Our study showed that
transpiration of water decreased by 61% and 85% after the shoots had been exposed to 186
and 1329 µg m-3 Hg-contaminated air for 3 days, respectively (Wang and Greger,
unpublished). We also found that the lower part of the leaves started to wilt after 3 days of
exposure to 186 µg m-3 Hg-contaminated air and all leaves became dry and partly brown
in 1329 µg Hg m-3.
Plants may survive in heavy metal-contaminated environment by preventing metals
from entering into the cytoplasm and because of the mechanism to detoxify the metals
inside the cytoplasm. When metals are initially absorbed by the roots, parts of them are
trapped in the cell wall, which reduces the amount of metals entering the cytoplasm. At
this stage, Hg has a high affinity for negative charges of pectin substances, hemicellulose
and cellulose. Our result showed that the cell wall is the major Hg binding component of
plant tissue, i.e. about 80% of Hg located in roots was bound to the cell walls, which is
similar to what has been reported for Pisum sativum L. and Mentha spicata L. by Beauford
et al. (1977). Hg taken into the cytoplasm of the cells is generally attributable to the
sequestration of toxic ions in complexes. Glutathione-related phytochelatins (Rajesh et al.,
1996; Zenk, 1996) are the most dominant molecules found so far to sequester metal ions.
However, from our results there is no evidence that phytochelatins are responsible for Hg
tolerance in willow, because no phytochelatins were detected in either sensitive or tolerant
willow clones. This has also been found in the case of other heavy metals (Landberg and
Greger, 2004). Another possible mechanism to detoxify Hg in plants is to release the
accumulated Hg into the air. Expression of genes merA and merB, originating from
bacteria, can detoxify Hg in transgenic plants, which convert hazardous methyl-Hg and
Hg2+ to volatile elemental Hg (Hg0) which is then released to the air (Bizily et al., 1999,
2000; Rugh et al., 1996). However, this is not the case for native plants, as our results
showed that no release of Hg from the shoot to the air was found in any of the six
investigated species (Paper I). Therefore, other mechanisms in the cytoplasm must be
operative to explain the Hg tolerance in willow.
29
Phytoremediation of Hg
Phytoextraction
In phytoextraction, metal-tolerant plants with high metal accumulation and high biomass
production are preferably used. Our results showed a large variation among the six clones
of willow in their sensitivity to Hg (paper II). The tolerant clone Björn was used to study
the phytoextraction of Hg both in pots with aged Hg-spiked soil or industrial Hgcontaminated soil and in the field. Results showed that this willow clone could grow
successfully without significant measurable toxic effects except with 1mM KI addition
(Papers III and IV). The toxic effects found in the test with 1 mM KI addition was thought
to be mainly due to the toxicity of iodide to the plants (Paper IV). It suggests that selected
willow clones are able to tolerate Hg while being used for phytoextraction of such types of
aged Hg-contaminated soil.
A possible release of Hg into air by plants may contribute to air contamination when
using phytoextraction in practice. However, our study showed that plant leaves do not
release Hg into the air in any of the investigated plant species (Paper I). This suggests that
there is no consequent increase of Hg burden in the atmosphere by phytoextraction.
Willow roots accumulated Hg from aged industrial Hg-contaminated soil (Papers III,
IV), as shown earlier for other plant species (Lenka et al., 1992). The plants used for
phytoextraction must have an ability to efficiently accumulate metal via their roots. Our
studies showed that willow roots efficiently accumulated Hg in hydroponics, where they
could accumulate more than 300 µg Hg g-1DW from of 1 µM Hg(NO3)2 (200 µg Hg L-1)
within 4 hours (Fig. 8) and reduce the Hg concentration in Hg(NO3)2 solution from initial
1 µM to 0.05 µM after 3 days of cultivation. Moreover, willow could accumulate Hg by
more than 1000 µg g-1DW in its roots without significant toxic effects (Paper II).
However, Hg accumulation in willow grown in soil was much less efficient than that of
willow grown in hydroponics (Papers II, III, and IV). Other plant species, e.g., western
thistle with the highest Hg accumulation among plant species grown in Hg-contaminated
soil at the Bohus site, accumulated similar low levels of Hg as willow (Table 4). The low
accumulation of Hg in plants from soil was believed to be due to the low bioavailability of
Hg in the soil. Indeed, the results of the sequential extraction showed that Hg in soil was
mainly bound to residual organic matter (53%) and sulphides (43%), which remained
stable during the cultivation of willow.
The low bioavailability of Hg in contaminated soil is a restricting factor in
phytoextraction of Hg. Compared with chelating agents, e.g. EDTA, iodide is more
efficient in mobilizing Hg in soil, which mainly forms the soluble complex HgI42- with a
stability constant of 29.8 (Wasay et al., 1995). However, too high iodide concentrations
may be toxic to willow (Paper IV; Mackowiak and Grossl, 1999; Zhu et al., 2003).
Therefore, the iodide concentrations used to increase the bioavailability should be
tolerated by plants. Additions of up to 1 mM KI increased the Hg concentrations to about
5, 3 and 8 times, respectively, in the leaves, branches and roots (Paper IV).
30
The plants used for phytoextraction should have high translocation of accumulated
metals to an easily harvestable part of the plant, i.e. the shoot in the case of willow.
However, both hydroponics and soil studies showed that willow had a low translocation of
Hg to the shoots (Papers I–IV), and similar results were found in other plant species (Paper
I; Beauford et al., 1977; Godbold and Hütterman, 1988). Moreover, although iodide
addition could increase the amount of Hg extracted by plants from soil, it could not
improve the low translocation of Hg from the roots to the shoots (Paper IV). The low
translocation of Hg to plant shoots detected leads to a low efficiency of Hg removal from
the contaminated soil if plant shoots alone are harvested. Hence, Hg-accumulating roots
should also be harvested together with shoots, which is apparently not feasible in practice.
Therefore, it might not be realistic to use this plant for phytoextraction of Hg in practice,
even though iodide could enhance the phytoextraction efficiency.
To estimate the time required to remove all Hg from a Hg-contaminated soil by using
phytoextraction, model calculations were made based on the data from field trials and pot
tests in Paper IV (Table 6). The calculations show that extremely long time is needed to
clean up the Hg-contaminated soils if stem alone is harvested. Moreover, industrial Hgcontaminated soil needs longer time to be cleaned up than Hg-spiked soil. This is due to
the differences in bioavailability of Hg between the two kinds of soils. The soil used in the
pot test was 1-year-old Hg-spiked agricultural soil and well homogenised with relatively
higher Hg bioavailability than that of the aged-soil in the field trial. The soil for the field
trial was polluted with Hg more than 30 years ago and was extremely heterogeneous.
Furthermore, it probably contained large amount of sulphur, as sulphur was previously
used by the company to produce sulphuric acids. The long ageing effect and the high
concentrations of sulphur lead to the extremely low bioavailability of Hg in the soil,
because the bioavailable Hg decreased with time by leaching, bacterial volatilization and
formation of stable Hg complexes with the soil matrix, especially with sulphur.
Table 6 Estimation of the time required to remove all Hg from two kinds of Hg-contaminated
soils by phytoextraction, assuming that the metal taken up by plants is from the top 50 cm of soil
harvest
Biomass production
Kg(ha*yr)-1 §
Hg in plant
µg (g DW)-1
years
Industrial Hg-contaminated soil with
50 mg Hg kg-1DW †
stem
23000
0.46
23600
root
16000
27.6
574
One-year-old Hg-spiked soil with
50 mg Hg kg-1DW ‡
stem
23000
0.70
15500
root
16000
274
57
Soil
† Calculation is based on the data from field trial at the site of a chlor-alkali plant in the vicinity of
Gothenburg (Sweden) with 0.5mM KI addition (Paper IV).
‡ Calculation is based on the data from pot tests with 0.2mM KI addition (Paper IV).
§ Biomass production of stem is based on the data from Labrecque and Teodorescu (2003). The root
biomass was based on the root/stem biomass ratio in the hydroponics cultivation.
31
Phytostabilization
In order to reduce the bioavailability or mobility of heavy metals, the plants used for
phytostabilization preferably have efficient root-accumulation of available metals in the
soil, low translocation of metals to the shoots, and a large root system. Willow roots could
efficiently accumulate Hg in hydroponics and had high affinities for Hg (Table 2; Papers
I–IV). Hg binds roots so hard that washing with 20 mM EDTA (30 min) only removed
less than 2% of total Hg in roots (Wang and Greger, unpublished). Therefore, willow roots
grown in Hg-contaminated soil were able to accumulate Hg and reduce its bioavailability
in soil (Table 7; Paper III). The exchangeable Hg and the Hg bound to humic and fulvic
acids decreased in the rhizospheric soil, whereas the plant accumulation of Hg increased
with the cultivation time. The sum of the decrease of these two Hg fractions in soil after 76
days of cultivation was approximately equal to the amount of the Hg accumulated in
plants, which accounted for about 0.2 % of the total Hg in soil. Moreover, the low
translocation of Hg to the shoots detected makes willow useful for phytostabilization of
Hg-contaminated land, in which root systems trap the bioavailable Hg and reduce the
leakage of Hg from contaminated soils (Fig. 12). However, the Hg-accumulated root
tissues may die and become debris. Bacterial activities on debris of Hg-accumulated
tissues need to be taken into account in long term cultivation.
Table 7 Hg bioavailability in aged industrial Hg-contaminated soil assessed by 1M MgCl2
extraction prior to and after cultivation of willow for 32 and 76 days. n = 3, ±SE (Data from
Paper III)
Treatment
Hg bioavailability (µg kg-1 soil DW)
Soil at start
46.1 ± 1.1 a
Rhizospheric soil
Day 32
31.3 ± 2.4 b
Day 76
18.2 ± 1.5 c
a,b,c Different letters in each column denote significant difference at α ≤ 0.05.
Phytostabilization may also partly result from physical effects, as the vegetation cover
can promote physical stabilization of a substrate, especially on sloping ground. Willow has
a massive root system, which helps to bind the soil. In addition, transpiration of water by
the willow reduces the overall flow of water down through the soil, thus, helping to reduce
the amount of Hg that is transferred to ground- and surface waters (Fig. 12).
Foliage filtration
Our present study showed that willow leaves were able to continuously absorb Hg from
air, and Hg concentrations in leaves and branches increased with prolonged exposure time
(Fig. 11). Hence, on a global scale, vegetation may function as a foliage filtration of Hg in
32
the air. However, relatively few data have been published so far on air-vegetation
exchange. The amount of Hg removed from the atmosphere by vegetation regionally or
globally is virtually unknown.
In consideration of food safety, uptake of Hg in vegetation from air contributes to part
of the intake of Hg by humans. Furthermore, atmospheric deposition is considered to
dominate the Hg input to most soils and lakes in the boreal forest zone, which causes Hgcontamination of fish (Meili et al., 2003). Therefore, global efforts are needed to reduce
the emission of Hg into the atmosphere.
Fig. 12 Illustration of phytostabilization of Hg.
33
Conclusions
It is apparent from this work that phytoextraction of Hg is promising and can be used
without unwanted release of Hg via stomata. However, although the roots efficiently take
up Hg from solution, the translocation to the shoots, i.e. the harvestable parts is low. The
low bioavailability in soil is also a limiting factor for using this technique, and even
though iodide can increase the bioavailability in soil and thus the uptake of Hg by plants,
the phytoextraction capacity is not large enough for aged Hg-contaminated soils. We did
not find any high Hg-accumulators among either the selected common cultivated plant
species or plants growing naturally at the Hg-contaminated sites. Among the willow
clones, known to commonly take up various levels of heavy metals, no difference in Hg
accumulation and translocation was found. To the previous findings we can also add that
none of the plant species was found to be suitable for phytoextraction. Therefore, it may
be concluded that phytoextraction is not a realistic technique to remediate Hgcontaminated soils. Nonetheless, as plant roots are able to efficiently take up Hg from the
available Hg pool in soil and to accumulate Hg in roots, phytostabilization might be a
promising approach to remediate aged Hg-contaminated soils. In this process, the massive
plant root systems trap the bioavailable Hg and reduce the leakage of Hg from
contaminated soil.
Future perspectives
Large variations in Hg sensitivity were found among the willow clones tested. However,
the Hg tolerance mechanism still remains an open question, since PCs, the most dominant
metal-tolerant molecules in plant, was not found in willow (Paper II; Landberg and
Greger, 2004). Further investigations are needed to find out the mechanism leading to Hg
tolerance in willow plants.
Willow grown in Hg-contaminated soil decreased the Hg bioavailability (Paper III).
This was conducted in our study with one type of soil during a 2.5-month period.
Additional investigations are needed to reveal whether Hg phytostabilization is operative
in long term cultivation as well as in various types of soils.
As the low translocation of Hg to the shoots is a restricting factor for phytoextraction
of Hg-contaminated soil, there is a need to search for plants with high translocation of Hg
to the shoots. Increased translocation of Hg to the shoots by genetic modification might be
an alternative option in Hg-phytoextraction.
34
Acknowledgements
Doing this thesis has been made easier and much more enjoyable by many great people, so
I would like to say a big thank to…
♥ Associate Prof. Maria Greger, my supervisor – thank you for giving me the opportunity
to do a PhD study and the great support (and firm words when needed!) you have given
me.
♥ Prof. Lena Kautsky, my co-supervisor – thank you for your support and the valuable
comments on the manuscripts and this thesis.
♥ Prof. Birgitta Bergman – thank you for your enlightening comments to this thesis as
well as to my Licentiate thesis.
♥ Prof. Marianne Pedersén – I am indebted for your encouragement and your great
support in my Licentiate.
♥ All people at the Department of Botany, especially members of the “Plant-metal group”
– Tommy, AnnHelén, Åsa, Eva, Agneta, Clara, Johanna, Lisa – I am forever grateful for
your support, warm company and always being so nice to me.
♥ Tommy, my big brother – you always give me help when I had problem with AAS,
HPLC, construction of experimental equipment, …
♥ Clara – you did a lot in assisting in my experiments, in language checking of my thesis
and etc etc…, I greatly appreciate all you have ever done for me.
♥ Eva – thanks for your countless helps and such a nice mid-summer festival for my wife
and me in Sala with your parents.
♥ Dimitra and Liang – for going over my thesis in the last and checking the reference list
♥ Our great gardeners in Botan – Peter and Ingela – for help to mange my greenhouse
experiments and for the beautiful colour and blooming scene you bring to us.
♥ Hans Lind – thank you very much for helping me to construct experimental equipments
♥ My friends in Inorganic Chemistry Department – Tang Liqiu, Peng Hong, and Liu Jing
– for help me making the gold traps.
♥ I’ve not been alone in doing this research – a big thank to all my co-authors! Special
thanks Catherine Keller for your help and advice on soil chemistry via email.
♥ For collaboration of Hg speciation study, I thank my Spanish colleagues, Prof. Carmen
Camara, Yolanda Madrid-Albarran, Pilar Ximenez-Embun and M. Eva Moremo.
35
♥ All my colleagues in COLDREM – for exchanging knowledge and the warm
discussion.
Work is not all in life, also, I would show my great appreciation for all the friends I have
made during my years in Stockholm, my PhD would not have been this enjoyable if not
for all these nice people around me…
♥ Our innebandy team in Botan – Clara, Johanna, Dietmar, Martin, Mathias, Herman,
Christine, Ingvild, Mats, Per, Sofia, Patrik, Prof. Stanislaw Karpinski, … – I greatly
enjoyed this Swedish sport and the happy time with you.
♥ Our TaiChi team in Botan – Mercedes, Johan, Eva, Åsa, Frida, Pernilla, Dietmar,
Martin, Liang, Sara, Lotta, Karolina, Regina, Daniel, Anders, Ulla, ... – it was a really
relaxing time with a lot of fun with your company.
♥ Brita, my previous landlady – you offered me the first “home” in Sweden, I did miss the
discussion on various topics with you.
♥ Gun – your great interests in Chinese culture and large collection of Chinese stuffs
really impressed us. I still remember the cosy Christmas Eve at your home.
♥ Suzanne and Petter – thank you for offering me the first sailing experience in my life, it
was so exciting!
♥ Tang Bing, Zhao Wei, Sun Yi, Chen Yunying, Yang Qian, Xie Yi, Feng Quanhong,
Tang Liqiu, Peng Hong, Wu Jiang, Huang Zhen, Jia Wei, Zhu Shunwei, and Jiang Ying –
for playing table tennis, badminton, and for all the nice time we had together.
♥ My fishing friends – Sun Yi, Tang Bing, Zhen Kang, Martin, and Rehab – a lot of
pleasure with you.
♥ Li Xin, Wang Jue, Huang Fang, Huang Qinghai, Ran Liang, and Liu Jing – for all
joyful time and delicious dinners we had together.
♥ My Belgian colleagues and friends – Prof. Max Mergeay, Prof. Daniël van der Lelie,
and Cindy – you are always willing to help me as soon as I ask.
♥ My family – my mum, dad, brother, older sister and younger sister – for your nonstopping believe in me and all your support.
♥ At last but of course not least, my wife – Yan – for your endless support,
encouragement and love.
This work was financially supported by the Swedish Foundation for Strategic
Environmental Research (MISTRA) through the research programme ‘Soil remediation in
a cold climate’ (COLDREM) and Carl Tryggers Foundation.
36
References
Allard, B., Arsenie, I., 1991. Abiotic reduction of mercury by humic substances in aquatic system – an
important process for the mercury cycle. Water Air and Soil Pollution 56, 457–464.
Barber, J., Beauford, W., Shieh, Y.J., 1973. Some aspects of mercury uptake by plant, algal and bacterial
systems in relation to its biotransformation and volatilization. In: Miller, M. W., Clarkson, T. W.
(eds) Mercury, mercurials, and mercaptans. Charles D Thomas Publisher, Illinois, USA. pp 325–345.
Barghigiani, C., Bauleo, R., 1992. Mining area environmental mercury assessment using Abies alba.
Bulletin of Environmental Contamination and Toxicology 49, 31–36.
Beauford, W., Barber, J., Barringer, A.R., 1977. Uptake and distribution of mercury within higher plants.
Physiologia Plantarum 39, 261–265.
Begley, T.P., Walts, A.E., Walsh, C.T., 1986. Mechanistic studies of a protonolytic organomercurial
cleaving enzyme – bacterial organomercurial lyase. Biochemistry 25 (22), 7192–7200.
Benoit, J.M., Gilmour, C.C., Mason, R.P., Heyes, A., 1999. Sulphide controls on mercury speciation and
bioavailability to methylating bacteria in sediment pore waters. Environmental Science and
Technology 33 (6), 951–957.
Bersenyi, A., Fekete, S., Hullar, I., Kadar, I., Szilagyi, M., Glavits, R., Kulcsar, M., Mezes, M., Zoldag,
L., 1999. Study of the soil-plant (carrot)-animal cycle of nutritive and hazardous minerals in a rabbit
model. Acta Veterinaria Hungarica 47 (2), 181–190.
Bizily, S.P., Rugh, C.L., Meagher, R.B., 2000. Phytodetoxification of hazardous organomercurials by
genetically engineered plants. Nature Biotechnology 18, 213–217.
Bizily, S.P., Rugh, C.L., Summers, A.O., Meagher, R.B., 1999. Phytoremediation of methylmercury
pollution: merB expression in Arabidopsis thaliana confers resistance to organomercurials.
Proceeding of the National Academy of Science of the United States of America 96, 6808¯6813.
Bloom, N.S., Crecelius, E.A., 1983. Determination of Mercury in Seawater at Sub-nanogram per Liter
Levels. Marine Chemistry 14, 49–59.
Bloom, N.S., Fitzgerald, W.F., 1988. Determination of volatile mercury species at the picogram level by
low-temperature gas-chromatography with cold-vapour atomic fluorescence detection. Analytica
Chimica Acta 208 (1-2), 151–161.
Bloom, N.S., Watras, J.J., 1988. Observations on Methylmercury in precipitation. Proceedings of
Conference on Trace Metal in Lakes, Hamilton. The Science of the Total Environment 87/88, 199–207.
Braeckman, B., Cornelis, R., Rzeznik, U., Raes, H., 1998. Uptake of HgCl2 and MeHgCl in an insect cell
line (Aedes albopictus C6/36). Environmental Research 79 (1), 33–40.
Browne, C.L., Fang, S.C., 1978. Uptake of mercury vapor by wheat. Plant Physiology 61, 430–433.
Cavallini, A., Natali, L., Durante, M., Maserti, B., 1999. Mercury uptake, distribution and DNA affinity in
durum wheat (Triticum durum Desf.) plants. The Science of the Total Environment 243/244, 119–127.
Clarkson, T.W., 1972. The pharmacology of mercury compounds. Annual Review of Pharmacology and
Toxicology 12, 375–406.
Compeau, G.C., Bartha, R., 1985. Sulfate-reducing bacteria: principal methylators of mercury in anoxic
estuarine sediment. Applied and Environmental Microbiology 50 (2), 498–502.
Coquery, M., Welbourn, P.M., 1994. Mercury uptake from contaminated water and sediment by the
rooted and submerged aquatic macrophyte Eriocaulon septangulare. Archives of Environmental
Contamination and Toxicology 26 (3), 335–341.
Cunningham, S.D., Berti W.R., 1993. Remediation of contaminated soils with green plants: an overview.
In Vitro Cellular and Developmental Biology 29, 207–212.
Cunningham, S.D., Ow, D.W., 1996. Promises and prospects of phytoremediation. Plant Physiology 110,
715–719.
37
Cunningham, S.D., Berti W.R., Huang J.W.W., 1995. Phytoremediation of contaminated soils. Trends
Biotechnology 13, 393–397.
D. Glass Associates Inc., 1999. U.S. and International Markets for Phytoremediation.
de Temmerman, L., Vandeputte, R., Guns, M., 1986. Biological monitoring and accumulation of airborne
mercury in vegetables. Environmental Pollution 41, 139–151.
Du, Sh.H., Fang, Sh.C., 1982. Uptake of elemental mercury vapor by C3 and C4 species. Environmental
and Experimental Botany 22 (4), 437–443.
Du, Sh.H., Fang, Sh.C., 1983. Catalase activity of C3 and C4 species and its relationship to mercury vapor
uptake. Environmental and Experimental Botany 23, 347–353.
Ebinghaus, R., Tripathi, R.M., Wallschläger, Lindberg S. E., 1999. Natural and anthropogenic mercury
sources and their impact on the air-surface exchange of mercury on regional and global scales. In:
Ebinghaus, R., Turner, R.R., Lacerda, L.D., Vasiliev, O., Salomons, W.(eds.) Mercury Contaminated
Sites: Characterization, Risk Assessment and Remediation. Springer - Verlag Berlin Heidelberg New
York. pp 1–50.
Ekvall, L., Greger, M., 2003. Effects of environmental biomass-producing factors on Cd uptake in two
Swedish ecotypes of Pinus sylvestris (L.). Environmental Pollution 121, 401–411.
Ensley, B.D., 2000 Rationale for use of phytoremediation. In: Raskin, I., Ensley, B.B. (eds.)
Phytoremediation of Toxic Metals: Using Plants to Clean Up the Environment. John Wiley & Sons,
Inc. pp 3–12.
Ericksen, J.A., Gustin, M.S., 2004. Foliar exchange of mercury as a function of soil and air mercury
concentrations. Science of the Total Environment 324 (1-3): 271–279.
Esau, K., 1953. Plant Anatomy. John Wiley & Sons, New York. p 493.
Fernández, J.A., Aboal, J.R., Carballeira, A., 2000. Use of native and transplanted mosses as
complementary techniques for biomonitoring mercury around an industrial facility. Science of the
Total Environment 256 (2-3), 151–161.
Fitzgerald, W.F., 1995. Is mercury increasing in the atmosphere – the need for an atmospheric mercury
network. Water, Air and Soil Pollution 80 (1-4), 245–254.
Fox, B., Walsh, C.T., 1982. Mercuric reductase. Purification and characterization of a transposon encoded flavoprotein containing an oxidation-reduction-active disulfide. Journal of Biological
Chemistry 257 (5), 2498–2503.
Godbold, D.L., 1991. Mercury induced root damage in spruce seedlings. Water Air and Soil Pollution 56,
823–831.
Godbold, D.L., 1994. Mercury in forest ecosystems: Risk and research needs. In: Watras, C. J., and
Huckabee, J. W., (eds) Mercury pollution – Integration and synthesis. Lewis Publishers, Boca Raton.
pp 295–303.
Godbold, D.L., Hütterman A., 1988. Inhibition of photosynthesis and transpiration in relation to mercuryinduced root damage in spruce seedlings. Physiologia Plantarum 74, 270–275.
Greger, M., Landberg, T., 1999. Use of willow in phytoextraction. International Journal of
Phytoremediation 1 (2), 115–123.
Greger M., Landberg T., Berg B., 2001. Salix clones with different properties to accumulate heavy metals
for production of biomass. Akademitryck AB, Edsbruk.
Hanson, P.J., Lindberg, S.E., Tabberer, T.A., Owens, J.G., Kim, K.H., 1995. Foliar exchange of mercury
vapor: evidence for a compensation point. Water, Air and Soil Pollution 80, 373–382.
Hempel, M., Thoeming J., 1999. Remediation techniques for Hg-contaminated sites. In: Ebinghaus, R.,
Turner, R.R., Lacerda, L.D., Vasiliev, O., Salomons, W. (eds.) Mercury Contaminated Sites:
Characterization, Risk Assessment and Remediation. Springer - Verlag Berlin Heidelberg New York.
pp 441–456.
38
Kalac, P., Svoboda, L., 2000. A review of trace element concentrations in edible mushrooms. Food
Chemistry 69 (3), 273–281.
Katz, S., Santilli, V., 1962. The reversible reaction of tobacco mosaic virus ribonucleic acid and mercury
chloride. International Journal of Biochemistry, Biophysics and Molecular Biology 55, 621–626.
Kawade, Y., 1963. The interaction of mercuric chloride with ribonucleic acids and polyribonucleotides.
Biochemical and Biophysical Research Communication 10, 204–208.
Kozuchowski, J., Johnson, D.L., 1978. Gaseous emissions of mercury from an aquatic vascular plant.
Nature 274, 468–469.
Labrecque, M., Teodorescu, T.I., 2003. High biomass yield achieved by Salix clones in SRIC following
two 3-year coppice rotations on abandoned farmland in southern Quebec, Canada. Biomass and
Bioenergy 25 (2), 135-146
Landberg, T., Greger, M., 1996. Differences in uptake and tolerance to heavy metals in Salix from
unpolluted and polluted areas. Applied Geochemistry 11, 175–180.
Landberg, T., Greger, M., 2004. No phytochelatin (PC2 and PC3) detected in Salix viminalis. Physiologia
Plantarum 121 (3), 481–487.
Lasat, M.M., 2002. Phytoextraction of toxic metals: A review of biological mechanisms. Journal of
Environmental Quality 31 (1), 109–120.
Laurier, F.J.G., Mason, R.P., Gill, G.A., Whalin L., 2004. Mercury distributions in the North Pacific
Ocean—20 years of observations. Marine Chemistry 90, 3–19.
Lenka, M., Panda, K.K., Panda, B.B., 1992. Monitoring and assessment of mercury pollution in the
vicinity of a chloralkali plant. 4. Bioconcentration of mercury in in-situ aquatic and terrestrial plants
at Ganjam, India. Archives of Environmental Contamination and Toxicology 22 (2), 195–202.
Lifvergren, T., 2001. Remediation of mercury polluted soil. PhD thesis, Örebro studies in Environmental
Science 1, Örebro University, Örebro, Sweden.
Lindberg, S.E., Stratton, W.J., 1998. Atmospheric mercury speciation: Concentrations and behaviour of
reactive gaseous mercury in ambient air. Environmental Science and Technology 32, 49–57.
Liu, Y., Cotgreave, I., Atzori, L., Grafstrom, R.C., 1992. The mechanism of Hg2+ toxicity in cultured
human oral fibroblasts: the involvement of cellular thiols. Chemico-Biological Interactions 85 (1),
69–78.
Lux, A., Šottníková, A., Opatrná, J., Greger, M., 2004. Differences in structure of adventitious roots in
Salix clones with contrasting characteristics of cadmium accumulation and sensitivity. Physiologia
Plantarum 120 (4), 537–545.
Macek, T., Mackova, M., Ka, J., 2000. Exploitation of plants for the removal of organics in
environmental remediation. Biotechnology Advances 18 (1), 23–34.
MacFarlane, G.R., Burchett, M.D., 2000. Cellular distribution of copper, lead and zinc in the grey mangrove,
Avicennia marina (Forsk.) Vierh. Aquatic Botany 68, 45–59.
Mackowiak, C.L., Grossl, P.R., 1999. Iodate and iodide effects on iodine uptake and partitioning in rice
(Oryza sativa L.) grown in solution culture. Plant and Soil 212, 135–143.
Marschner, H., 1995. Mineral Nutrition of Higher Plants. Academic Press, London.
Martian-Doimeadios, R.C.R., Wasserman, J.C., Bermejo, L.F.G., Amouroux, D., Nevado, J.J.B., Donard,
O.F.X., 2000. Chemical availability of mercury in stream sediments from the Almaden area, Spain.
Journal of Environmental Monitoring 2 (4), 360–366.
Matilainen, T., Verta, M., Korhonen, H., Uusi-Rauva, A., Niemi, M., 2001. Behaviour of mercury in soil
profiles: Impact of increased precipitation, acidity, and fertilization on mercury methylation. Water,
Air and Soil Pollution 125 (1-4), 105–119.
Mauseth, J.D., 1988. Plant Anatomy. The Benjamin/Cummings Publishing Co., Inc., Menlo Park CA. pp 277.
McGrath, S.P., Zhao, F.J., 2003. Phytoextraction of metals and metalloids from contaminated soils.
Current Opinion in Biotechnology 14 (3), 277–282.
39
Meili, M., Bishop, K., Bringmark, L., Johansson, K., Munthe, J., Sverdrup, H., de Vries, W., 2003.
Critical levels of atmospheric pollution: criteria and concepts for operational modelling of mercury in
forest and lake ecosystems. Science of the Total Environment 304 (1-3), 83–106.
Mierle, G., 1988. The determination of Mercury in Lakes, Streams and Precipitation. Proceeding
Conference Trace Metal in Lakes, Hamilton. The Science of the Total Environment 87/88, 299–302.
Munthe, J., Wängberg, I., Iverfeldt, A., Lindqvist, O., Stromberg, D., Sommar, J., Gardfeldt, K., Petersen,
G., Ebinghaus, R., Prestbo, E., Larjava, K., Siemens, V., 2003. Distribution of atmospheric mercury
species in Northern Europe: final results from the MOE project. Atmospheric Environment 37, S9-S20.
Ogata, M., Aikoh, H., 1984. Mechanism of metallic mercury oxidation in vitro by catalase and
peroxidase. Biochemical Pharmacology 33, 490–493.
Olson, B.H., Cooper, R.C., 1976. Comparison of aerobic and anaerobic methylation of mercuric chloride
by San Francisco Bay sediments. Water Research 10, 113–116.
Ortiz, A.I.C., Madrid-Albarran, Y., Camara-Rica, C., 2002. Evaluation of different sample pre-treatment
and extraction procedures for mercury speciation in fish samples. Journal of Analytical Atomic
Spectrometry 17 (12), 1595–1601.
Patra, M., Sharma, A., 2000. Mercury toxicity in plants. The Botanical Review 66 (3), 379–421.
Porcella, D.B., Chu, P., Allan, M.A., 1996. Inventory of North American mercury emissions to the
atmosphere: relationship to the global mercury cycle. In: Baeyens, W., Ebinghaus, R., Vasiliev, O. (eds.)
Global and regional mercury cycles: sources, fluxes and mass balances. NATO-ASI Series 2.
Environment vol 21. Kluwer, Dordrecht, The Netherlands. pp 179–190.
Pulford, I.D., Watson, C., 2003. Phytoremediation of heavy metal-contaminated land by trees – a review.
Environment International 29 (4), 529–540.
Rajesh, K.M., Jose, M.V., Ramana, K., Rizwana, A., Thomas, C.H., Priti, M., 1996. Optical spectroscopic
and reverse-phase HPLC analyses of Hg(II) binding to phytochelatins. Biochemical Journal 314, 73–82.
Raskin, I., Kumar, P.B.A.N., Dushenkov, S., Salt, D.E., 1994. Bioconcentration of heavy metals by
plants. Current Opinion in Biotechnology 5, 285–290.
Raskin, I., Smith, R.D., Salt, D.E., 1997. Phytoremediation of metals: using plants to remove pollutants
from the environment. Current Opinion in Biotechnology 8, 221–226.
Reeves, R.D., Baker, A.L.M., 2000. Metal-accumulating plants. In: Raskin, I., Ensley, B.B. (eds.)
Phytoremediation of Toxic Metals: Using Plants to Clean Up the Environment. John Wiley & Sons,
Inc. pp 193–229.
Ribeyre, F., Boudou, A., 1994. Experimental study of inorganic and methylmercury bioaccumulation by
four species of freshwater rooted macrophytes from water and sediment contamination sources.
Ecotoxicology and Environmental Safety 28, 270–286.
Rugh, C.L., Senecoff, J.F., Meagher, R.B., Merkle, S.A., 1998. Development of transgenic yellow poplar
for mercury phytoremediation. Nature Biotechnology 16, 925–928.
Rugh, C.L., Wilde, H.D., Stack, N., Thompson, D.M., Summers, A.O., Meagher, R.B., 1996. Mercuric
ion reduction and resistance in transgenic Arabidopsis thaliana plants expressing a modified bacterial
merA gene. Proceedings of the National Academy of Sciences of the United States of America 93,
3182–3187.
Salt, D.E., Blaylock, M., Kumar, N.P.B.A., Viatcheslav, D., Ensley, B.D., 1995. Phytoremediation: a
novel strategy for the removal of toxic metals from the environment using plants. Biotechnology 13,
468–474.
Salt, D.E., Smith, R.D., Raskin, I., 1998. Phytoremediation. Annual Review of Plant Physiology and Plant
Molecular Biology 49, 643¯668.
Schiering, N., Kabsch, W., Moore, M.J., Distefano, M.D., Walsh, C.T., Pai, E.F., 1991. Structure of the
detoxification catalyst mercuric ion reductase from Bacillus sp strain-rc607. Nature 352, 168–172.
40
Schuster, E., 1991. The behavior of mercury in the soil with special emphasis on complexation and
adsorption processes – a review of the literature. Water, Air and Soil Pollution 56, 653–666.
Siegel, S.M., Puerner, N.J., Speitel, T.W., 1974. Release of volatile mercury from vascular plants.
Physiologia Plantarum 32, 174–176.
Slemr, F., Langer, E., 1992. Increase in global atmospheric concentrations of mercury inferred from
measurements over the Atlantic Ocean. Nature 355 (6359), 434–437.
Steinnes, E., 1997. Mercury. In: Alloway, B.J. (ed.) Heavy metals in soils (second edition). Blackie
Academic & Professional Press, London, UK. pp 245–259.
Stratton, W.J., Lindberg, S.E., 1995. Use of a refluxing mist chamber for measurement of gas-phase
water-soluble mercury (II) species in the atmosphere. Water, Air and Soil Pollution 80, 1269–1278.
Suhadra, A.V., Panda, K.K., Panda, B.B., 1993. Residual Mercury in seed of Barley of methanesulfate,
maleic hydrazide, methyl mercury chloride and mercury-contaminated soil. Mutation Research 300
(34), 141–149.
Taylor, G.J., Stadt, K.J., Dale, M.R.T., 1991. Modelling the interactive effects of aluminium, cadmium
manganese, nickel and zinc stress using the Weibull frequency distribution. Canadian Journal of
Botany 69 (2), 359–367.
Tchounwou, P.B., Ayensu, W.K., Ninashvili, N., Sutton, D., 2003. Environmental exposure to mercury
and its toxicopathologic implications for public health. Environmental Toxicology 18 (3), 149–175.
Ullrich, S.M., Tanton, T.W., Abdrashitova, S.A., 2001. Mercury in the aquatic environment: A review of
factors affecting methylation. Critical Reviews in Environmental Science and Technology 31 (3), 241–293.
von Burg, R., Greenwood, M.R., 1991. Mercury. In: Merian E. (ed.) Metals and Their Compounds in the
Environment VCH Verlagsgesellschaft mbH, Weinheim.
Wagner–Dobler, I., von Canstein, H., Li, Y., Timmis, K.N., Deckwer, W.D., 2000. Removal of mercury
from chemical wastewater by microoganisms in technical scale. Environmental Science and
Technology 34, 4628–4634.
Wang, Y.D., Greger, M., Ximenez-Embun, P., Madrid-Albarran, Y., Camara-Rica, C., 2002 Mercury
speciation in phytoextraction of mercury. In: Mench, M., Mocquot, B. (eds.) COST Action 837 WG2
Workshop 2002 – Risk assessment and sustainable land management using plants in trace elementcontaminated soil, Bordeaux, France. p 56.
Wängberg, I., Edner, H., Ferrara, R., Lanzillotta, E., Munthe, J., Sommar, J., Sjoholm, M., Svanberg, S.,
Weibring, P., 2003. Atmospheric mercury near a chlor-alkali plant in Sweden. Science of the Total
Environment 304 (1-3), 29–41.
Wasay, S.A., Arnfalk, P., Tokunaga, S., 1995. Remediation of a soil polluted by mercury with acidic
potassium–iodide. Journal of Hazardous Materials 44, 93–102.
WHO, 1976. Mercury. Environmental Health Criteria 1. World Health Organization, Geneva. pp 1–132
WHO Regional office for Europe, 2000. Air quality guidelines – Second edition, Copenhagen, Denmark.
Chapter 6.9. pp 1–15.
Xiao, Z.F., Stromberg, D., Lindqvist, O., 1995. Influence of humic substances on photolysis of divalent
mercury in aqueous-solution. Water Air and Soil Pollution 80 (1-4), 789–798.
Zenk, M.H., 1996. Heavy metal detoxification in higher plants – A review. Gene 179 (1), 21–30.
Zhu, Y.G., Huang, Y.Z., Hu, Y., Liu, Y.X., 2003. Iodine uptake by spinach (Spinacia oleracea L.) plants
grown in solution culture: effects of iodine species and solution concentrations. Environment
International 29, 33–37.
Österås, A.H., Ekvall, L., Greger, M., 2000. Sensitivity to and accumulation of Cadmium in Betula
pendula, Picea abies, and Pinus sylvestris seedlings from different regions in Sweden. Canadian
Journal of Botany 78, 1440–1449.
41