Phytostabilisation - use of wetland plants to treat mine tailings

Phytostabilisation  -  use of wetland plants to treat mine tailings
- use of wetland plants to treat
mine tailings
Doctoral thesis
Eva Stoltz
Department of Botany
Stockholm University, 2004
© Eva Stoltz
ISBN 91-7265-972-6 pp. 1-45
PrintCenter, Stockholm University
Sweden 2004
Phytostabilisation – use of wetland plants to treat mine tailings
Mine tailings can be rich in sulphide minerals and may form acid mine drainage (AMD) through
reaction with atmospheric oxygen and water. AMD contains elevated levels of metals and
arsenic (As) that could be harmful to animals and plants. An oxygen-consuming layer of organic
material and plants on top of water-covered tailings would probably reduce oxygen penetration
into the tailings and thus reduce the formation of AMD. However, wetland plants have the
ability to release oxygen through the roots and could thereby increase the solubility of metals
and As. These elements are released into the drainage water, taken up and accumulated in the
plant roots, or translocated to the shoots.
The aim was to examine the effects of plant establishment on water-covered mine tailings
by answering following questions: A) Is plant establishment on water-covered mine tailings
possible? B) What are the metal and As uptake and translocation properties of these plants? C)
How do plants affect metal and As release from mine tailings, and which are the mechanisms
Carex rostrata Stokes, Eriophorum angustifolium Honck., E. scheuchzeri Hoppe,
Phragmites australis (Cav.) Steud., Salix phylicifolia L. and S. borealis Fr. were used as test
plants. Influences of plants on the release of As, Cd, Cu, Pb, Zn and in some cases Fe in the
drainage water, and plant element uptake were studied in greenhouse experiments and in the
The results obtained demonstrate that plant establishment are possible on water-covered
unweathered mine tailings, and a suitable amendment was found to be sewage sludge. On
acidic, weathered tailings, a pH increasing substance such as ashes should be added to improve
plant establishment. The metal and As concentrations of the plant tissue were found to be
generally higher in roots than in shoots. The uptake was dependent on the metal and As
concentrations of the tailings and the release of organic acids from plant roots may have
influenced the uptake. The metal release from tailings into the drainage water caused by E.
angustifolium was found to depend greatly on the age and chemical properties of the tailings.
However, no effects of E. angustifolium on As release was found. Water from old sulphide-,
metal- and As-rich tailings with low buffering capacity were positively affected by E.
angustifolium by causing higher pH and lower metal concentrations. In tailings with relatively
low sulphide, metal and As contents combined with a low buffering capacity, plants had the
opposite impact, i.e. a reduction in pH and elevated metal levels of the drainage water. The total
release of metal and As from the tailings, i.e. drainage water together with the contents in shoots
and roots, was found to be similar for C. rostrata, E. angustifolium and P. australis, except for
Fe and As, where the release was highest for P. australis. The differences in metal and As
release from mine tailings were mainly found to be due to the release of O2 from the roots,
which changes the redox potential. Release of organic acids from the roots slightly decreased
the pH, although did not have any particular influence on the release of metal and As.
In conclusion, as shown here, phytostabilisation may be a successful technique for
remediation of mine tailings with high element and sulphide levels, and low buffering capacity.
Doctoral dissertation, 10 Dec, 2004
Eva Stoltz
Department of Botany
Stockholm University, Sweden
© Eva Stoltz
ISBN 91-7265-972-6 pp. 1-45
PrintCenter, Stockholm University
Sweden 2004
Table of contents
Abstract .........................................................................3
Table of contents..............................................................4
1 Introduction..................................................................6
1.1 Mine waste ................................................................................. 6
1.2 Remediation techniques ................................................................. 8
1.3 Phytoremediation ......................................................................... 9
1.3.1 Plant metal uptake and translocation .......................................... 11
1.3.2 Consequences of oxygen translocation by wetland plants .................. 11
1.3.3 Root exudates ....................................................................... 11
1.4 Summary of the introduction .......................................................... 13
1.5 Aim ......................................................................................... 14
2 Comments on materials and methods................................. 15
Mine sites.................................................................................. 15
The choice of plant species ............................................................ 15
The choice of amendments............................................................. 15
Germination of seeds.................................................................... 16
Quality control – metal analyses ...................................................... 17
Experimental design in papers III and IV............................................. 17
Calculations in the thesis............................................................... 17
Attempt to study the buffering effect of plants in paper II ..................... 17
Description of tailings in an additional experiment............................... 18
3 Results and discussion ................................................... 19
3.1 Is plant establishment on water-covered mine tailings possible?............... 19
3.1.1 Suitable amendments.............................................................. 19
3.1.2 Plant species that may tolerate different types of stress................... 21
3.2 What are the metal and As uptake and translocation properties in wetland
plants growing on water-covered mine tailings?........................................ 22
3.2.1 Factors affecting the metal and As plant uptake............................. 22
3.2.2 Translocation properties .......................................................... 25
3.2.3 Comparing metal and As concentrations in plant shoots .................... 27
3.3 How do plants affect metal and As release from mine tailings, and which are
the mechanisms involved?................................................................... 28
3.3.1 pH effect on the drainage water from mine tailings by Eriophorum
species....................................................................................... 28
3.3.2 Observations explaining the pH effect of E. angustifolium................. 30
3.3.3 Influences of E. angustifolium on the release of metals and As from the
drainage water in tailings with different mineral composition ................... 33
3.3.4 Amount of metals and As that could be dispersed by E. angustifolium .. 34
3.3.5 Influences of different plant species on the total release of metals and As
from mine tailings rich in sulphides .................................................... 36
3.3.6 Metal and As release from mine tailings - long-term perspective? ........ 37
4 Conclusions................................................................. 39
Acknowledgements.......................................................... 40
References.................................................................... 41
List of papers
This thesis is based on the following papers, which will be referred to by their
Roman numerals:
Stoltz, E. and Greger, M. 2002. Accumulation properties of As, Cd,
Cu, Pb and Zn by four wetland plant species growing on submerged
mine tailings. Environ. Exp. Bot. 47: 271-280.
Stoltz, E. and Greger, M. 2002. Cottongrass Effects of Trace
Elements in Submersed Mine Tailings. J. Environ. Qual. 31: 14771483.
Stoltz, E. and Greger, M. Effects of different wetland plant species
on sulphidic mine tailings. Submitted to Plant and Soil.
Stoltz, E. and Greger, M. The release of metal and As from various
mine tailings by Eriophorum angustifolium. Submitted to Applied
Stoltz, E. and Greger, M. Influences of wetland plants on weathered
acidic mine tailings. Submitted to Environmental Pollution.
My contributions to the papers were as follows: I was responsible for
the writing of all papers with help from the co-author. I planned the
experiments in all papers with help and advice from the co-author. I
performed most of the laboratory work in all papers, and also the
collection of field material in paper V.
Reprints of papers I and II were made with permission from
Elsevier (I) and ASS, CSSA, SSSA (II).
1 Introduction
1.1 Mine waste
Waste from mines causes major environmental problems all over the world.
Heavy metals and As contents in mine waste are much higher than in
uncontaminated soil. Some elements that are commonly found in high
concentrations in mine waste are As, Cd, Cu, Pb and Zn. Metal and As
concentrations in different tailings and in uncontaminated soil are shown in
Table 1. All of these elements are, in elevated levels, toxic to animals and
plants (McDowell, 1992).
Table 1. Total heavy metal content in different mine wastes
soils (µg (gDW)-1)
Uncontaminated soil, Sweden
22-0.7 100 200 10-300
Uncontaminated soil, Ireland
and in uncontaminated
Borgegård and Rydin
Milton and Johnson
Kristineberg mine site, Sweden
Unlimed tailings
Limed tailings
Boliden mine site, Sweden
Tailings with vegetation
Fresh tailings
1069 1063
914 1133
1424 2044 14464 paper I
849 1972 3861 papers III and IV
Garpenberg mine site, Sweden
paper IV
Aitik mine site, Sweden
paper IV
Laisvall mine site, Sweden
paper IV
Copper tailings from Laver,
Åmmeberg Zinc mine, Sweden
Holmström et al. (1999)
Bergholm and Steen
Ye et al. (1999)
Lead/Zinc mine, Guandon,
Copper mine in West Cornwall, 3010
paper II
paper II
In Sweden, a total of more than 600 Mtonnes of mine wastes have been
deposited (MiMi, 2001). The amount of waste produced annually in Sweden
from sulphide and iron ores is 45 Mtonnes, out of which 20 Mtonnes are mine
tailings and the rest is waste rock (MiMi, 2001). Mine tailings are a finegrained sand material and the remains after the desired metals have been
extracted. The tailings are mixed with water and flushed through pipelines out
into large impoundments for deposition (Fig. 1).
Figure 1. Tailings impoundment at the Boliden mine site, Northern
Many ores that are rich in metals consist of pyrite (FeS2) and other minerals
rich in sulphides. The oxidation of sulphides generates protons, metals and As,
hence, acid mine drainage (AMD) is produced (Lowson, 1982; Salmon, 2003).
The oxidation agents could be atmospheric oxygen or ferric iron (Fe3+), which
is produced by further oxidation of the ferrous iron (Fe2+) formed by the
oxidation of pyrite. Table 2 shows some proton-generating and/or metal
releasing reactions by sulphide weathering. In addition, the same table shows
proton-consuming reactions by the weathering of, e.g., carbonate minerals such
as calcite, and silicate minerals such as chlorite, plagioclase and potassium
feldspar, minerals that may also be present in mine tailings. Thus, if acid mine
drainage will be formed or not depends on the mineral composition of the
tailings. Also, weathering rates of the different minerals have an impact. The
weathering rates of the various minerals are in general: silicate minerals <
sulphides < carbonate (Strömberg and Eriksson, 1996).
Table 2. Proton generating and proton consuming reactions by weathering of
minerals that may occur in mine tailings
Proton producing and metal releasing reactions
Oxygen path
FeS2 + H2O + 7/2O2 → Fe2+ + 2SO 4 + 2H+
CuFeS2 + 4O2 → Fe2+ + Cu2+ + 2SO 4
ZnS + 2O2 → Zn2+ + SO 4
PbS + 2O2 → Pb2+ + SO 4
FeAsS + 3.25O2 + 1.5H2O → Fe2+ + HAsO 4 + SO 4 + 2H+
Ferric iron path
FeS2 + 14Fe3+ + 8H2O → 15Fe2+ + 2SO 4 + 16H+
CuFeS2 + 16Fe3+ + 8H2O → 17Fe2+ + Cu2+ + 2SO 4 + 16H+
ZnS + 8Fe3+ + 4H2O → 8Fe2+ + Zn2+ + SO 4 + 8H+
PbS + 8Fe3+ + 4H2O → 8Fe2+ + Pb2+ + SO 4 + 8H+
Proton consuming reactions
CaCO3 + 2H+ → Ca2+ + CO2 + H2O
(Mg4.5Fe 0.2 Fe 0.2 Al)AlSi3O10(OH)8 + 16H+ → 4.5Mg2+ + 0.2Fe2+ +
0.2Fe3+ + 2Al3+ +3SiO2 + 12H2O
Na0.75Ca0.25Al1.25Si2.75O8 + 5H+ → 0.75Na+ + 0.25Ca2+ + 1.25Al3+ +
2.75SiO2 + 2.5H2O
KAlSi3O8 + H+ + 4H2O → K+ +3H4SiO4 + Al(OH)3
From Strömberg and Eriksson (1996), Holmström (2000), Carlsson (2002) and Salmon (2003).
1.2 Remediation techniques
Different techniques for AMD treatment have been studied, e.g. metal
precipitation by creating an anoxic environment (Cheong et al., 1998) and
water treatment by constructed wetlands (Sobolewski, 1999). Additionally,
several methods to prevent the formation of AMD by reducing the contact of
water and/or atmospheric oxygen with the sulphide mineral have been
investigated. Nyavor et al. (1996) summarise chemical, biochemical and
physical processes for this purpose.
Two common physical techniques are dry-cover and water-cover. The dry
cover often consists of a dense material, e.g., fine moraine, placed on top of the
tailings to reduce the penetration of atmospheric oxygen into the tailings and in
this way creating an anoxic environment. The dry-cover may also consist of
two layers, one is a dense sealing layer of fine compact material to prevent
oxygen diffusion, and the other is a cover layer of e.g. moraine, for erosion
protection (Clemensson-Lindell et al., 1992; Bergström, 1997; Elander et al.,
The water-cover, or flooding, consists of a high water table (1.5-2 m) over
the tailings to reduce wave erosion and resuspension of the tailings as well as
oxygen transportation into the tailings (Elander et al., 1998). As a consequence,
the oxidation of the sulphides in water-covered tailings is reduced. In addition,
lime is often added to the water in the last impoundment located before the
water outlet for precipitation of dissolved metals. However, a high water table
requires high impoundment walls that are expensive to build, and also the
pressure from the water might reduce the stability of the walls (Grimalt et al.,
1999). Therefore, phytoremediation of the metals should be considered on
water-covered tailings, in order to reduce the water level and thereby reduce the
costs and risks of high impoundment walls.
1.3 Phytoremediation
Phytoremediation is defined as the use of green plants to remove, contain, or
render environmental contaminants harmless (Cunningham and Berti, 1993).
Phytoremediation of heavy metals can be divided into three groups:
phytoextraction – metal accumulating plants are established on contaminated
soil and later harvested in order to remove metals from the soil; rhizofiltration
– roots of metal accumulating plants absorb metals from polluted effluents and
are later harvested to diminish the metals in the effluent; and phytostabilisation
– metal tolerant plants are used to reduce the mobility of metals, thus, the
metals are stabilised in the substrate (Salt et al., 1995). One reason to use plants
for remediation concerns the relatively low cost and maintenance requirements
(Cunningham and Berti, 1993).
Among the above-mentioned phytoremediation techniques, phytostabilisation might be the most appropriate one to use for remediation of mine
tailings. Phytostabilisation may be a successful way to prevent the formation of
AMD and stabilise the metals in the tailings (Tordoff et al., 2000). Plants with
low shoot accumulation should be used in order to stabilise the metals and As
in the tailings and reduce the metal dispersion through grazing animals or at
leaf senescence. When plants are established on water-covered mine tailings,
an organic layer will eventually be formed on top of the mine waste (Fig. 2).
This layer would probably consume oxygen (due to the chemical and
biological processes) and together with the plants also reduce wind and wave
erosion. As a benefit, a high water table and impoundment walls are not
Tailings containing
Organic layer
Figure 2. Plant establishment on water-covered tailings might stop the
penetration of oxygen.
Since mine tailings are poor in plant nutrient content, an amendment to
improve conditions for plant establishment should be added (Borgegård and
Rydin, 1989; Jung, 2001; Ye et al., 2001). As amendment, waste products of
organic material should be considered for low costs, also contributing to the
formation of the organic, oxygen-consuming layer. However, a vegetation
cover of the submersed tailings requires wetland plants, which are known to be
able to release oxygen through their roots (Armstrong, 1992). This mechanism
could oxygenate the tailings and increase formation of AMD (Fig. 3).
Furthermore, plant roots release exudates, i.e. organic acids, protons and CO2,
which could increase the weathering of certain minerals (Kelly et al., 1998).
Organic layer
O2, CO2, organic
acid, H+, CO 3 ,
Tailings containing
OH- release from
plant roots?
Me+ and As plant
uptake and
translocation to shoot?
Me+ and As leakage?
Figure 3. Possible effects of an organic layer and plants growing on
submerged mine tailings. Me+= metal ions.
1.3.1 Plant metal uptake and translocation
Most heavy metals and As are toxic to plants, although at low levels some of
the metals are essential for plant growth (Marschner, 1995). Above a certain
level also these metals become toxic to the plant, yet plant roots may still take
up such elements (Marschner, 1995). Some species have evolved a tolerance to
heavy metals enabling them to grow in contaminated soils (Baker, 1987; Baker
and Walker, 1990). In addition, some wetland species seem to have an inherent
tolerance to heavy metals that does not have to be developed (Wu, 1990;
McCabe and Otte, 2000; Matthews et al., 2004). Many plants that are tolerant
to metals and As accumulate the elements in the root tissue and have a low
translocation to the shoots, hence, the [element]shoot /[element] root ratio is low
(Coughtrey and Martin, 1978; Baker and Walker, 1990). Plants with these
uptake and translocation properties are suitable to use for establishment of
mine tailings, consequently reducing the dispersion of metal and As.
1.3.2 Consequences of oxygen translocation by wetland plants
When plant roots grow, oxygen is required for the respiration process. Many
plant species living in wetlands or other aquatic systems are adapted to the
anaerobic environment and translocate gas from the atmosphere to the
underground organs through a lacunar system of intercellular airspaces or
through aerenchyma, (Armstrong, 1992; Brix, 1993). Oxygen produced in the
photosynthesis may also be used (Chen and Barko, 1988). The oxygen released
into the water-covered mine tailing deposits might mobilise the metals and As,
which, in reduced conditions, are bound to sulphides (Table 2). Thereby,
making metals and As available for plant uptake or leakage into the
environment with the drainage water (Fig. 3).
1.3.3 Root exudates
Plants may alter the pH conditions in soils, which have a major impact on the
weathering of some minerals (Drever, 1994; Kelly et al., 1998; Hisinger et al.,
2001). Also, organic acids that may be released from plant roots (Jones and
Darrah, 1994; Marschner, 1995) have been found to increase the release of
metals from mine tailings (Fig. 3) (Burckhard et al., 1995; Wasay et al., 2001).
The release of CO2 from root respiration may reduce pH of the tailings
(Kelly et al., 1998). Furthermore, pH changes by plants may be due to the
imbalance in cation or anion release, caused by the excess of either anion or
cation uptake (Haynes, 1990; Marschner, 1995). Such a situation could occur
when, for example, the nitrogen source is nitrate, causing pH to increase due to
the fact that plants exchange nitrate with, e.g., OH- or CO 32− ions (Nye, 1981;
Marschner and Römheld, 1983). When the nitrogen source is ammonium,
protons will be used for the exchange and the pH will decrease (Fig. 4)
(Marschner and Römheld, 1983; Villegas and Fortin, 2001).
The release of low-molecular-weight organic acids (e.g. citric acid, malic
acid) is also believed to be an important factor for the weathering of minerals
(Ochs, 1996). Organic acids may cause pH changes, but the effects suggested
in different studies are contradictory. Some studies show that a decrease in pH
is given by organic acids (Hoffland, 1989; Marschner, 1995; Kelly et al.,
1998). In other studies, it has been suggested that organic acids could increase
the pH, since the pH of plant cells is most commonly ca 7.2, and in this
environment the organic acids should be present as anions. Thereby, the
organic acids could take up protons instead of releasing them when entering the
soil. Still, a pH increase requires a pH of the substrate lower than 7.2, and the
counterbalancing cation during passage of the organic acid across the cell
membrane cannot be H+ (Fig. 4) (Jones and Darrah, 1994; Jones, 1998).
Therefore, the impact on pH from the release of organic acids is not clear.
Imbalance in cation/anion release
Ammonium uptake
pH decrease
NH 4
Plant root cells
Nitrogen uptake
pH increase
NO 3
Organic acids (OA) release
pH < 7.2
pH 7.2
nK , nNa
No pH change or pH
pH increase
nK , nNa
pH decrease
Figure 4. Release of substances from roots that may affect soil pH. The
mechanisms are discussed in: Nye (1981), Marschner and Römheld (1983),
Hoffland (1989), Jones and Darrah (1994), Marschner (1995), Villegas and
Fortin (2001).
1.4 Summary of the introduction
Even though mine tailings are low in nutrients, high in metal and As
concentrations, as well as being acidic if weathered, plant establishment is
probably possible if suitable amendments are added. It is more difficult to
conclude if phytostabilisation is a suitable remediation technique for watercovered mine tailings from the information in the introduction. The organic
layer that will be formed on top of the tailings will most likely reduce O2
penetration. However, since wetland plants are known to release O2 by the
roots, it is difficult to predict if the final condition of the tailings will be
reduced or aerated. Furthermore, plants have several other mechanisms that
could release elements from the minerals, e.g. by exuding H+, organic acids or
other root exudates. The metal and As uptake and translocation properties of
the plants, to reduce the dispersion of elements via the shoots, must be
1.5 Aim
The aim of the present study was to examine the role of plant establishment on
metal and As release from water-covered mine tailings by trying to find the
answers to following questions:
1. Is plant establishment on water-covered mine tailings possible?
2. What are the metal and As uptake and translocation properties in
wetland plants growing on water-covered mine tailings?
3. How do plants affect metal and As release from mine tailings, and
which are the mechanisms involved?
2 Comments on materials and methods
2.1 Mine sites
Five mine sites in Sweden were used for the
studies in this thesis (Fig. 5). The main site was
Boliden mine area (64º 52’ N, 20º 22’ E), where
most of the material was collected (I; III; IV; V)
and where the field experiments (I; V) were
performed. Tailings from the Kristineberg mine
area (65º 04’ N, 18º 44’ E) were used in paper
II. In paper IV, apart from Boliden, the tailings
were collected from the mine areas at Aitik (67º
04’ N, 19º 09’ E), Laisvall (66º 14’ N, 17º 12’
E) and Garpenberg (60º 19 N, 16º 09’ E).
2.2 The choice of plant species
The plant species used were bottle sedge (Carex
rostrata) (I; III; V), common cottongrass
(Eriophorum angustifolium) (I; II; III; IV; V),
white cottongrass (E. scheuchzeri) (II), common
reed (Phragmites australis) (I; III; V), and two
willow species (Salix phylicifolia and S.
borealis) (I). All plant species were common
wetland species at either the Boliden or the Figure 5. The locations of
some mine sites in Sweden.
Kristineberg mine sites. The two Salix species
that were used in paper I showed shoot
accumulation properties of Cd and Zn, which is not unexpected since Salix is
known to have high shoot concentrations of these metals (Brieger et al., 1992;
Landberg and Greger, 1996). Since species with high shoot element
accumulation are not suitable for establishments of mine tailings, those species
were not used in the further studies. Furthermore, E. scheuchzeri was only
studied in paper II, since it showed similar effects as E. angustifolium when
growing on mine tailings, apart from that it did not tolerate the high pH that
was found for E. angustifolium.
2.3 The choice of amendments
In a pre-study to paper II, growths of Eriophorum angustifolium and
Phragmites australis in different waste products added to unlimed mine
tailings from Boliden mine area were tested to find out the most suitable
amendments for plant establishment. The ashes used were bio-ashes from a
heating plant using wood chips as fuel. The bio sludge and green liquor dregs
are waste products from the pulp and paper industry, the former is a product
from the mechanical pulp industries and the latter is produced in sulphate pulp
production (Greger et al., 1998). The peat was a peat used for plantation
(Rölunda produkter AB, Bålsta, Sweden). The results showed that controls
with no amendment addition had no, or very poor, plant growth (Table 3).
Sewage sludge and the ash-sewage sludge mixture gave the greatest plant
growth (Table 3), which was why those two amendments were used in paper II.
Table 3. Comparison of establishment of Eriophorum angustifolium and
Phragmites australis in different amendments. A=Ashes, SS=sewage sludge,
P=Peat, BS=Bio sludge, GLD=Green liquor dregs, n = 6
GLD Control SS
Pots with plant establish83
100 50
ment out of 6 (%)
Total dry weight in all
25.1 0.5
pots (g)
Mean dry weight of the 6
0.10 1.66
0.45 0.13
0.05 4.18 0.13
pots (g)
The reason why only sewage sludge was used in papers III and IV was that the
addition of only sewage sludge gave the largest biomass of E. angustifolium in
both limed and unlimed, and of E. scheuchzeri in unlimed, unweathered
tailings in paper II.
In paper V both sewage sludge and ashes-sewage sludge mixture were used
since the pH enhancing effect of ashes (Greger et al., 1998) was believed to be
a positive influence for plant growth in the acidic environment of weathered
2.4 Germination of seeds
Seeds from the wetland plants used in the experiments were found to be
difficult to germinate if they were not pre-treated correctly. If the seeds were
not used directly after collection, they were stored wet and cold (4°C). If the
seeds were stored longer than one month, they germinated when placed in
warm and light conditions without any extra treatment. If the seeds were not
stored wet and cold, they had to be treated with diurnal fluctuation in both
temperature and light, i.e., 12-h dark period with 6 °C, and a 12-h light period
of 19 °C over at least 5 days, to be able to germinate. The light fluctuation
might not be necessary since diurnal fluctuation of temperature alone has been
found to stimulate the germination of many wetland species (Thompson et al.,
2.5 Quality control – metal analyses
As reference material, reed canary grass (Phalaris arundinaceae L, Reference
Material NJV 94-4,) was used for grasses and sedges in all papers, and Willow
(Salix, Reference Material NJV 94-3) for the two Salix species in paper I. Both
reference materials were obtained from the Swedish University of Agricultural
Sciences, Uppsala, Sweden.
For all metal (Cd, Cu, Fe, Pb and Zn) and As analyses where the atomic
absorption spectrophotometer (AAS-100) was used, either by flame, furnace or
vapour generation, the standard addition technique was used to eliminate the
matrix effect.
2.6 Experimental design in papers III and IV
The experiments in papers III and IV were performed at the same time with the
same methods, therefore, the results on Boliden tailings with control, sewage
sludge and sewage sludge + E. angustifolium treatments were used in both of
the studies.
2.7 Calculations in the thesis
In section 3.3.3, a comparison between the released metals and As from Aitik
(IV) and Kristineberg (II) tailings is made. Since there were differences in
element-accumulation time of the drainage water in the two studies, the data
was adjusted. The element accumulation time was 6.5 times shorter in paper II
(2 months) compared with that in paper IV (13 months), and the element release
was assumed to be proportional; the element content of the drainage water in
paper II was multiplied by 6.5. Furthermore, the amount of the drainage water
from Kristineberg tailings (II) was not measured, but in the calculations
assumed to be 0.1 L, similar to the volume of the free space where the drainage
water was collected. The same adjustment was made when the amount of metals
and As released that might affect the environment (element content in shoot +
drainage water) was calculated and compared between papers II and IV, as
shown in Fig. 11 (see 3.3.4).
2.8 Attempt to study the buffering effect of plants in paper II
Plants had an increasing and stabilising effect on drainage water pH from
sulphide-rich tailings, as described in paper II. An attempt to find the plant
mechanism behind the pH increase was made by measuring organic acids and
alkalinity of the drainage water from tailings in the studies of papers III and IV.
However, the mechanisms involved could not be investigated since the tailings
had high buffering capacity, except for the Aitik tailings, and no increase in pH
or alkalinity by plants was found (III; IV). Plants reduced the alkalinity and the
pH in drainage water from Aitik tailings, but the buffering mechanism by
plants, as seen in paper II, was not found. Thus, the buffering capacity of plants
could not be studied in any of the tailings.
2.9 Description of tailings in an additional experiment
The chemical properties of the Boliden
mine tailings used in the additional
greenhouse experiment discussed in
section 3.3.2 are shown in Table 4.
Table 4. Chemical properties of the
Boliden mine tailings
Fe *103
(mg kg-1)
2923± 470
29.5± 1.6
2268± 267
277± 22
2432± 420
6880± 2120
ICP-AES, SGAB Analytica, n=1
AAS, n=6, menas±SE
3 Results and discussion
3.1 Is plant establishment on water-covered mine tailings
3.1.1 Suitable amendments
Plant establishment on mine tailings is possible if an amendment is added
(Borgegård and Rydin, 1989; Ye et al., 2001; II; III; IV; V). The amendment
that seemed to be suitable on many kinds of different unweathered tailings was
sewage sludge (Table 5), which resulted in good plant establishment
(Borgegård and Rydin, 1989; II; III; IV). The addition of sewage sludge
resulted in higher plant biomass compared with the addition of ashes-sewage
sludge mixture in paper II, both in unlimed and lime treated tailings. The
reason for the poor plant growth in the ashes-sewage sludge mixture might
have been the initial impacts of the ashes. Ashes have the ability to increase the
pH (Greger et al., 1998), which then might have been too high for the seedlings
added in paper II, especially in the lime treated tailings. In addition, when
ashes and sewage sludge are mixed, the high pH of the ashes may cause
nitrogen loss through formation of volatile ammonia (Greger et al., 1998).
Moreover, the ashes-sewage sludge treatment only contained two-thirds of the
sewage sludge (with high N levels) compared with the treatment with only
sewage sludge (II). Furthermore, the very high Ca content in ashes (Greger et
al., 1998) can bind to phosphorus (Schnoor, 1996), making it less available to
plants. Hence, the level of available nutrients for the plants decreases in the
ashes-sewage sludge mixture.
On weathered acidic tailings the addition of the ashes-sewage sludge
mixture resulted in higher plant biomass, one year after start of the experiment,
than if only sewage sludge had been added (V) (Table 5). In addition, two
years after the start of the experiment, no plants had survived with only sewage
sludge addition (V). Thus, the positive effect of only sewage sludge addition
found in unweathered tailings was not seen in weathered acidic tailings. This
was probably not due to the acidity directly, since many wetland plants can
grow in an acidic environment (Nixdorf et al., 2001). Instead, the reason might
be that the acidity creates an environment with high levels of soluble metals
that might have a direct toxic effect on the plant growth (Baker, 1987; Baker
and Walker, 1990; Marschner, 1995). It could also be due to an indirect
toxicity created by high concentrations of free metals and As ions in the water
that could be taken up by the plants instead of nutrients, by competition
(Greger et al., 1991; Marschner, 1995). Furthermore, the latter
Table 5. Number of replicates with successful establishment of E. angustifolium
out of the total number (in brackets), and plant biomass per dm-2 growing in
different tailings with various amendments in greenhouse and field experiments.
SS= sewage sludge, A=ashes, nd=not determined
Type Thickness
(g dm-2)
Unweathered unlimed tailings
6 (6)
4.9 ± 0.60
6.6 ± 0.9
3 (6)
1.3 ± 0.02
2.1 ± 0.2
6 (6)
17.7 ± 1.00
11.8 ± 1.0
6 (6)
28.1 ± 2.60
14.7 ± 2.8
6 (6)
21.4 ± 1.70
15.8 ± 0.9
6 (6)
17.6 ± 1.00
8.8 ± 1.0
Unweathered limed tailings
5 (6)
3.2 ± 0.40
2.7 ± 0.4
0 (6)
Weathered tailings
5 (5) 0.005 ± 0.001
5 (5) 1.140 ± 0.320
effect may be enhanced by a dry period when the water levels are low and the
metals and As concentrations increase even more. The addition of ashes
increases the pH (Greger et al., 1998), which may precipitate the metals that are
otherwise soluble in acidic tailings (Schnoor, 1996). Thus, the metal
concentrations are reduced and the environment becomes less harmful for the
plants (V).
The high biomass of plants in paper IV may be due to the greater amount of
amendment added compared with that in paper II (Table 5). The biomass of
plants grown in the field experiment cannot be compared with the greenhouse
experiment, since it did not consist of the total shoot biomass (V). The shoots
of the plants in the field were cut off 0.1 m above the ground to ensure that the
plants survived to the next season for further studies. By the time the plants
were cut off they were about 0.4 m high in sewage sludge + ashes amendment
and the shoots were shorter (ca 0.2 m) in only sewage sludge amendment;
therefore a relatively large part of the shoots was left in the field.
In unweathered tailings without amendment the plant growth was very poor
(III; IV; V). This is probably due to the lack of nutrients, since the plants grew
when sewage sludge was added. On the other hand, the pH was around 7 and
the free metals and As concentrations were low (Stoltz, unpublished data), thus
the growth conditions for plants, except for the nutrients, were relatively good.
In other studies, it has been shown that plant establishment is possible without
the addition of amendment (Wright and Otte, 1999; McCabe and Otte, 2000;
Ye et al., 2001; Jacob and Otte, 2004a). This must depend on the chemical
properties of the tailings and if the tailings are naturally fertilized by, e.g.,
surface water or rain containing nutrients. However, plant growth was often
increased by fertilizer addition (Ye et al., 2001; Jacob and Otte, 2004a).
3.1.2 Plant species that may tolerate different types of stress
Many wetland plant species are able to cope with the high metal and As
concentrations of mine tailings (Coombes, 1997; McCabe and Otte, 1999;
Wilkinson et al., 1999; Jacob and Otte, 2004; I; II; III; IV; V). Eriophorum
angustifolium has been found to be able to survive in substrates with a wide pH
range, from pH 10.9 in paper II to about 2.7 in paper V. The lowest pH that E.
angustifolium has been found to grow in under field conditions is 2.6 (Nixdorf
et al., 2001). Other wetland plant species that have been found to grow on mine
tailings and are known to tolerate a low pH are Carex rostrata, E. scheuchzeri,
Phragmites australis, Typha angustifolia, T. latifolia (Jacob and Otte, 2004;
Wilkinson et al., 1999; I; II; III; V). These species have been found growing
under field conditions in pH as low as 2.1, 4.4, 2.1, 3.0 and 2.5, for each
species, respectively (Nixdorf et al., 2001).
However, it is not certain that wetland plant species can survive in an
environment with both high metals and As levels and low pH. In paper V,
plants in the field experiment did not survive after two years in weathered
tailings with low pH and probably high available metal and As concentrations,
even though nutrients in the form of sewage sludge were added. Still, in a
greenhouse experiment, plants survived in the same treatment as in the field
(V). In addition, there were naturally established plants at the field site of paper
V that survived in the tailings without any known nutrient supply. The
successful plant growth in the greenhouse experiment might be due to the
relatively short period of time they were studied (8 months), since in the field,
plants survived after one year of the experiment but not after two years.
Another reason for the growth of the plants in the greenhouse experiments
might have been due to the regular watering of the plants, which kept the
metals and As concentration of the pore water constant. In the field, there were
dryer periods that made the metals and As levels of the water higher, which
might have been too toxic for the plants to survive in. The reason why the
naturally established plants were able to survive might have been due to those
wetland plants species being spread by rhizomes, and many plants are
connected to each other. Thereby, if there is a spot with good conditions for the
plants, e.g., high nutrient levels, either on the surface or at a certain depth in the
tailings, the plants might be able to transport the nutrients from the high
nutrient area to plants in areas with lower nutrient levels.
3.2 What are the metal and As uptake and translocation
properties in wetland plants growing on water-covered mine
3.2.1 Factors affecting the metal and As plant uptake
Plant metal and As uptake may depend on different factors, e.g. external
conditions such as element concentrations of the substrate, and in what form
the metals and As are present. Plant mechanisms that influence the metal and
As uptake could be root exudates, e.g. organic acids, CO2 and H+ that change
the pH, which may release the elements from the substrate (Kelly et al., 1998).
In addition, the ability of the plant to take up or avoid uptake of the chelated or
free ions has an impact (Baker and Walker, 1990). Furthermore, other
compounds that are released from roots such as sugars may not have a direct
impact on the metal release or uptake by plants. Nonetheless, sugars and also
amino acids and organic acids exuded from roots may be a nutrient source for
bacteria, which might influence the release of metals and As from the soil and
the plant uptake and tolerance of the elements (Drever 1994; Marschner, 1995;
Burd et al., 2000; Glick, 2003).
Some positive correlations between the concentrations of metal and As in
plant tissue and the organic acid concentration of the drainage water were
found in paper III. The correlations were only found within the different plant
species, and not between them (III). Therefore, the release of organic acids
may influence the concentrations of metals and As both in roots and shoots,
although differently for the various species. These differences between the
various species might be due to different uptake mechanisms among the
species. Even though there were positive correlations in all species, i.e. the
greater the release of organic acids, the higher the tissue concentrations of
metals and As, the plant species might differ in rate and ability to take up the
metals and As as complexes or as ions. Other studies have also shown that
organic acids may have impact on the plant tissue metals and As concentrations
(Harmens et al., 1994; Ma et al., 2001).
The growth rate also influences the element tissue concentrations; in a plant
species with high growth rate the metal and As concentrations will probably be
diluted. Different species may vary in the efficiency to pick up a certain
element that they need. For example, a species that tolerates high levels of an
element has mechanisms to either avoid uptake, e.g., excluders with low tissue
concentrations, or the ability to make the element harmless within the plants,
e.g. accumulators with high tissue concentrations (Baker, 1987; Baker and
Walker, 1990), compared with species that do not tolerate such conditions. For
the case in paper III, all species seemed to tolerate the growth conditions. Still,
there might have been small variations in uptake of the elements that caused
the differences in impact of organic acids. Also, there were some differences in
biomass among the plants that might influence the metals and As
concentrations, since elements may be diluted by greater biomass (III).
The shoot and root metal and As concentrations of E. angustifolium in
papers II and IV were correlated with the total metal and As concentrations of
the tailings in Figure 6. The shoot concentrations of Cu and Pb showed a large
increase when the tailings concentrations were above 900 and 6000 mg kg-1,
respectively (Fig. 6). Similar results were found in roots for those two elements
and also for Cd and Zn. A great increase in root concentration was observed
when the Cd, Cu, Pb and Zn concentrations of the tailings exceeded 30, 900,
1500 and 9000 mg kg-1, respectively. Thus, E. angustifolium seems to have a
protection mechanism that prevents Cd, Cu, Pb and Zn uptake by the roots up
to a certain level of the substrate. This mechanism is suggested to be
‘avoidance of uptake’ that has been found in other plant species (Baker, 1987;
Baker and Walker, 1990). The low uptake of those metals could be due to the
formation of Fe plaque (Fe oxides or hydroxides) on the roots of wetland plants
by the release of O2. When large amounts of Fe are deposited on the roots,
metals may be adsorbed to the plaque and thereby reduce the metal uptake into
the roots (Otte et al., 1989; Greipsson and Crowder, 1992).
The translocation of Cu and Pb to the shoots seemed to depend on the root
concentrations (Fig. 6), whereas the translocation of Cd and Zn must be
dependent on other factors. The translocation was about 10 times lower for Pb
than for Cd and Cu, and 50 times lower than that of Zn.
The concentrations of As in roots increased proportionally with the
concentrations of the tailings. Hence, the plant root As concentration reflected
the external conditions by some uptake mechanism; this type of uptake is
element concentration in shoot (mg kg )
element concentration in root (mg kg )
element concentration in tailings (mg kg )
Figure 6. The relationship between the metal and As concentrations in
various mine tailings and in shoots and roots of E. angustifolium. Data from
paper II, ○= unlimed Kristineberg tailings △= limed Kristineberg tailings and
paper IV, ■= Boliden tailings, ▲= Garpenberg tailings, ◆= Laisvall tailings
and ●= Aitik tailings.
called ‘indicator plant’ by Baker and Walker (1990). The same pattern was
found for Zn shoot concentrations, which did not depend on the root
concentrations, therefore the shoot Zn concentrations may be used as an
indicator of the Zn concentrations in the tailings. The translocation of As was
low, about 50 times lower than that of Zn and similar to that of Pb, and it did
not seem to depend on the root concentrations. Thus, the mechanism to protect
the plant from As toxicity was probably low translocation to the shoot, possibly
by complex binding with phytochelatins and storage of the element in the root
(Sneller, 1999). Zinc may also be complex bound to, e.g., organic acids and
thereby become less harmful to the plants (Thurman and Rankin, 1982;
Harmens et al., 1994).
3.2.2 Translocation properties
Many species that are able to grow on mine tailings have been found to have
the highest metal and As concentrations in the roots, with a few exceptions
(Coombes, 1997; Ye et al., 2001; I; II; III; IV). There may be a mechanism that
protects the photosynthetic plant parts from toxic levels of metals and As
(Coughtrey and Martin, 1978; Landberg and Greger, 1996). One way to
tolerate high metal levels is the binding of metals to the cell walls of the roots
(Ernst et al., 1992; Marschner, 1995). In a crude cell wall extraction made on
roots of E. angustifolium and C. rostrata that had been grown in hydroponics
containing Hoagland nutrient solution and 0.5 µM Cd for 2 weeks, 71 % of the
Cd content was found in the cell walls of E. angustifolium (Greger et al.,
unpublished results). For C. rostrata, there were no significant differences
between the Cd content in the root and the crude wall extraction (Greger et al.,
unpublished results). Metals may also be chelated and could be transported to
the cell vacuole by, e.g., phytochelatins, organic acids, and amino acids for
storage (Ernst et al., 1992; Clemens, 2001).
Species with low translocation of metals and As to the shoots are suitable
for vegetation establishment on mine tailings to reduce the element dispersion
via grazing animals or at leaf senescence. However, not only the translocation
properties of the plants should be taken into consideration when planning to
vegetate mine tailings, but also the actual shoot metals and As concentrations,
since that may be toxic to grazing animals even though the species have been
found to be root accumulators (I). Furthermore, the opposite results have also
been found (IV), where E. angustifolium grown in tailings with relatively low
metal and As concentrations showed accumulation properties for Cd, Cu and
Zn both in shoot and root. Even though the plant had accumulating properties,
the concentrations in the shoots of those metals were lower or similar to the
tolerable levels of metals for animals (NCR, 1980). In addition, the Cd, Cu and
Zn concentrations of the shoots in plants with accumulation properties were
lower or similar to the concentrations of E. angustifolium grown in other
tailings with higher metal and As levels (IV) that did not show accumulation
properties. Thus, a plant that shows shoot-accumulating properties might
anyway be suitable for plant establishment on mine tailings, when looking at
the actual concentrations of the shoot.
When studying the metal and As uptake and translocation in plants, it is
important to do so in the appropriate substrate. Plants grown in a greenhouse
experiment with hydroponics, containing the same metal and As concentrations
as the possible plant extractable fraction of the tailings, did not have similar
uptake and translocation as in plants actually growing in the tailings in the field
(I). This may be due to the processes in the soil-root interphase that are not
present in hydroponics (I). In hydroponics, most metals are most likely present
as cations and attracted to the negatively charged cell walls of the roots,
therefore, they may be taken up easily by the plants (Fig. 7a). Arsenic is
suggested to be taken up as the anion of arsenate (As(V)) with mechanics
similar to that of phosphorus (Otte and Ernst, 1994 ). In soil or in this case,
mine tailings, the metals and As are bound within the mineral or on clay or
organic particles (from the sewage sludge), and some part of the metals and As
may be present as ions, which may make the plant uptake different to that in
hydroponics (Fig. 7b). Arsenic may be included in the minerals or precipitated
on Fe oxides or hydroxides (Sracek, 2004), or be present as anions.
Furthermore, there are interactions in the tailings between roots and bacteria
that are not present in hydroponics, which may impact the metals and As
uptake (Burd et al., 2000).
The possible plant extractable fraction may not be the actual fraction that
the plant takes up. Thus, the concentrations of metal and As ions in the
hydroponics might be higher or lower than the concentrations that the plants
are able to release in the tailings. Differences in the climate conditions, i.e., in
the growth cabinet and in the field, may also have an impact on the metals and
As uptake (Greger, 1999).
The number of negatively charged sites, i.e. the cation exchange capacity
(CEC) on tailings particles or roots, ‘compete’ for the metals, and since
different species may vary in their root CEC, the element uptake is affected
(Marschner, 1995). Different cultivation methods may also affect distribution
of metals and As within the plants. Carex rostrata grown in tailings had a
significantly higher shoot:root ratio for Cd, Cu and Pb, compared with
hydroponically grown plants (I). The same result was found for Salix and Cd
-- - Men+
- AsAs
- Men+
- As
-Men+- -
-- As
- - - - As- Men+ - - Men+
n+- Me
AsMe - Men+ - - n+
- Me
- As- Me-n+
- -n+ - -
Figure 7. Interactions between metals and As and roots in a) hydroponics and
b) mine tailings.
3.2.3 Comparing metal and As concentrations in plant shoots
The shoot metals and As concentration of the plants growing in mine tailings is
important to study since, together with the drainage water, it is a source for
element dispersion into the environment. The metals and As in the roots are
probably not released, instead they will remain in the roots and rhizomes as
long as these remain alive - C. rostrata, E. angustifolium and P. australis are
perennial species. When the roots die, the metals and As will most likely
become a part of the material in the upper part of the tailings, either bound to
organic material or, if the conditions are reduced, precipitate as sulphides
(Schnoor, 1996). The element concentrations in shoots of E. angustifolium
from plants growing on mine tailings were 1-2.6, 1-2.5, 1-3.6, 1-7.3 and 1-5.0
times higher for As, Cd, Cu, Pb and Zn, respectively, compared with the
concentrations of a ‘normal’ plant (Table 6). The metal and As concentrations
of the shoots in plants growing in sewage sludge in the field (V) were higher
than those presented in Table 6, but since the plant growth was poor and they
did not survive more than one season they were not included. Even though the
metal and As shoot concentrations were in some cases higher than that of a
‘normal’ plant, the levels were much lower than in hyperaccumulator plants
(Table 6). Hence, E. angustifolium is almost certainly suitable for plant
establishment on mine tailings since the possible metals and As dispersion
from the shoots was only slightly higher than that of a ‘normal’ plant, even
though growing in mine tailings.
Table 6. Metal and As concentrations in: 1) shoots of Eriophorum angustifolium
grown in different mine tailings, 2) ‘normal’ plants and 3) the lowest levels
required for a plant to be called a hyperaccumulating plant
(mg kg-1)
E. angustifolium in tailings
0.7-2.0 7.3-53.3 4.0-72.8 187-1018
‘Normal plant'
Hyperaccumulating plant
From I; II; IV; V 2 From Pais and Jones, 2000 3 From Baker et al.,2000;
Raskin and Ensley, 2000
Some differences in plant shoot concentrations were found between E.
angustifolium, C. rostrata and P. australis growing in the same type of tailings
(I; III). However, in most cases, the differences in shoot concentrations were
only between 1-4 times. The greatest differences were found in paper I, where
E. angustifolium had, respectively, 9.3 and 8.4 times higher Cu and As
concentrations than P. australis. Generally there were no significant
differences between the species investigated in dispersion of metals and As via
the leaves, therefore, from this point of view all of the species are probably
suitable for plant establishment on mine tailings (I; III).
3.3 How do plants affect metal and As release from mine
tailings, and which are the mechanisms involved?
3.3.1 pH effect on the drainage water from mine tailings by
Eriophorum species
Plants have been found to increase the sulphate and in some case the O2 levels
of the drainage water (II; III; IV), and in addition also increase the redox
potential in water-covered mine tailings (Wright and Otte, 1999; Jacob and
Otte, 2004a; III; IV; V), except for P. australis in paper III. Wetland plants are
known to be able to translocate O2 to the roots for root respiration, and the
increase in redox potential, sulphate and O2 concentrations may be due to some
of the O2 being released from the roots (Armstrong, 1992; Brix 1993).
Furthermore, wetland plants may change the redox condition to detoxify
reduced substances in the soil or facilitate the uptake of nutrients (Chen and
Barko, 1988; Laan et al., 1989). The release of oxygen would increase the
weathering of sulphides and thereby increase the metal and As content and
reduce the pH of the drainage water. However, in paper II, the metal levels
were reduced by two Eriophorum species in sulphide-rich unlimed tailings, and
pH was around 5.5 at the end of the experiment (Tables 3, 4 in II), while
treatments without plants had higher metal concentrations and a pH around 3.
The plants did not affect the concentration of As of the drainage water. The
relationship between metal and As concentrations and pH in drainage water
from paper II is shown in Figure 8.
The unlimed tailings used in paper II must have had a low buffering
capacity since the pH dropped from 5 to less than 3 in controls within 6 months
(Fig. 9c). In the lime-treated tailings studied in paper II and in other studies,
plants showed a slight pH decrease compared with controls, but the pH was
relatively high, above 6.2 in all studies (Wright and Otte, 1999; Jacob and Otte,
2004a; Jacob and Otte 2004b; III; VI). Also, in some of those cases, plants
Control Pots with
Pots with plants
and amendments
Cd (µg ml )
Cu and Pb (µg ml )
Pots with
Pots with plants
and amendments
As (µg l )
Zn (µg ml )
Figures 8 a and b. The relationship between pH and a) Cu, Pb and Cd b) Zn
and As concentrations in drainage water from unlimed tailings. Areas
between the two lines in the diagram show which treatment the values
originate from in paper II.
increased the metal concentrations in the drainage water. The tailings of those
studies must have had high buffering capacity or low content of sulphides,
since the pH of the drainage water from treatments without plants did not
decrease during the time of the experiment. The observed pH decrease by
plants could have either been due to increase sulphide weathering by root O2
release ore to the release of organic acids. A negative correlations between the
pH and the concentration of citric acid were found in paper III for all of the
plant species used, i.e., C. rostrata, E. angustifolium and P. australis. Thus, in
this study, organic acids caused a pH-reduction and not a pH-increase that was
discussed in section 1.3.3. Since no pH decrease was found in treatments
without plants, the buffering effect of plant found in paper II could never be
observed in papers III and IV. In weathered, acidic, tailings in paper V, E.
angustifolium did not have an influence on the pH. This may be due to poor
plant viability because they only survived when the pH was elevated by the
addition of amendment containing ashes that probably reduced the
concentrations of released metals.
3.3.2 Observations explaining the pH effect of E. angustifolium
Three observations in different sulphide-rich tailings were made that might
explain the high pH levels in the drainage water from pots with E.
angustifolium at the end of the experiment in paper II, compared with
treatments without plants. Firstly, those in the greenhouse experiment in paper
V with unweathered tailings (Fig. 9a and Fig. 1 in V), secondly, those found in
another tailings collected at the Boliden mine site, with a start pH of 7.6, with
which the same greenhouse experiment was performed as in paper V (Fig. 9b)
(see 2.9 for description of the tailings), and thirdly, the results of an additional
pH measurement performed after four months of the experiment but not
included in paper II (Fig. 9c).
In all three observations mentioned above (Fig. 9), plants reduced the pH in
water from unweathered tailings initially, but then the plants had the ability to
increase the pH. The reason that the treatments without plants did not decrease
in pH in the two Boliden tailings (Figs 9a and b) was most likely the high
buffering capacity of the tailings. The initial reduction in pH was possibly due
to plants increasing the weathering of sulphides by the release of O2, and then
the plants must have been able to sense the pH reduction and have had the
ability to counteract the process. However, since E. angustifolium is known to
be able to live in acidic environments (see 3.1.2), perhaps it was not the pH
reduction as such the plant responded to, instead it might be the increased
SS + E. a
SS + E. s
Figure 9 a, b and c. Effect of plants at different times on the pH of the
drainage water from pots with a) Boliden tailings used in paper V, b) Boliden
tailings with high buffering capacity in a separate experiment performed in a
greenhouse using the same treatments as in paper V, c) Kristineberg tailings
with low buffering capacity from paper II (see 3.3.2 for details). SS=sewage
sludge, E. a= Eriophorum angustifolium, E. s= E. scheuchzeri.
metal concentrations of the water due to the weathering and reduction in pH
(Fig. 10, Step 1). Hence, to prevent toxic metal levels in the pore water, the
plants were not able to reduce the weathering process but could increase the pH
that reduces the metal concentrations of the drainage water (Fig. 10, Step 2).
The increase in pH could be achieved by release of hydroxide ions and
carbonate ions, also the release of organic acids could possibly cause a pH
increase (see 1.3.3).
Those results could also explain why plants did not survive in weathered
acidic tailings in paper V at low pH, i.e., probably due to high metal and As
concentrations of the pore water, which are commonly found in water of
weathered tailings (Jung, 2001; II). When an amendment that increased the pH
was used, the metal concentrations of the water were almost certainly reduced
and plant establishment was successful. However, in this case the plants
themselves did not have the ability to increase the pH (V). A stabilising effect
of the pH by plants could also be seen in paper III and IV. The pH of drainage
water from pots with plants was relatively stable during the experiment
compared with the other treatments, except in Aitik tailings (Fig. 2 in IV).
The suggestion that E. angustifolium increased the pH in paper II to reduce
toxic metal levels of the water was tested in a pilot study. Seeds of the plant
species were found to germinate in acidic water both with high and low metal
and As concentrations, however, two weeks after germination, all germinated
seeds had died in the former treatment but not in the latter. Therefore, too high
metal levels in acidic pore water were found to be toxic to E. angustifolium.
Step 1
O2 + sulphides
SO 4
SO 4
Step 2
O2 + sulphides
CO 3 ?
Figure 10. Influences of plants on mine tailings rich in sulphides. See 3.3.2 for
3.3.3 Influences of E. angustifolium on the release of metals and As
from the drainage water in tailings with different mineral composition
Different effects of E. angustifolium on metal release in the drainage water
from various mine tailings were found in paper IV. However, the plant did not
affect the As concentrations of drainage water in any of the tailings, possibly
because the chemical properties of As are different to those of metals; for
example, arsenic forms anions and not cations, as well as arsenic being more
mobile in reduced than in aerated conditions (Masscheleyn et al., 1991; Sracek
et al., 2004). It is also possible that the roots took up all the arsenic released
from the tailings by the plants, and consequently plants did not cause any
increase of As in the drainage water. In that case, plants were unable to take up
As released by other factors, since no decrease was found either.
The highest metal and As release to the drainage water in relation to the
concentrations in the tailings was found in the tailings with the lowest total
metal and As concentrations (IV). Thus, the metal and As release cannot be
predicted from the amount of metals and As in the tailings. This is probably
due to the differences in mineral composition and chemical properties of the
tailings, such as buffering capacity, sulphide content and weathering rate
(Tables 1, 2 in IV).
In tailings with high sulphide levels and high buffering capacity (Boliden
tailings, VI), the sulphide weathering will eventually exceed the proton
consuming reaction and AMD will be produced. On those types of tailings,
plant establishment will probably eventually show a similar positive influence
as observed in paper II, with a metal release reduction and an increase in pH of
the drainage water.
In tailings with very low levels of sulphides (Laisvall tailings, IV), plants
might not have any particular impact on the metal release and pH. Even though
a slight pH decrease was found in the Laisvall tailings with plants, it was not so
great that it would have a major effect on the solubility of the metals. In
tailings with very high buffering capacity (Garpenberg tailings, IV), the proton
consuming minerals will most likely exceed the proton producing reaction,
which could be seen in a very stable pH (Fig. 2 in IV). Thus, plants will not
show an increased release of metals to the drainage water, due to the high pH.
The Aitik tailings in paper IV, with relatively low sulphide levels and low
buffering capacity of the drainage water, did not show a pH increase by plants
as in paper II. Eriophorum angustifolium resulted in a continuous pH decrease
and a high release of Cu and Fe in the drainage water, due to the weathering of
chalcopyrite (Table 2; IV). The low levels of total metal and As release could
explain why plants did not show a buffering effect in Aitik tailings. If E.
angustifolium responded to increased levels of free metals instead of the
reduced pH in paper II, as discussed in 3.3.2, the released metals from the Aitik
tailings were not too high for the plants to cope with. Even though the pH was
greatly reduced, the accumulated total content of metals (Cd, Cu, Pb and Zn)
and As in the drainage water during the experiments from Aitik tailings was
much lower, 218 µg (0.8 mg L-1), than that of Kristineberg tailings with plant
treatment, 3771 µg (7.5 mg L-1) (see 2.7 for information about the
calculations). This suggestion was confirmed by the pilot study mentioned in
3.3.4 Amount of metals and As that could be dispersed by E.
In Figure 11, the metals and As content in the drainage water, together with
shoot content, is displayed, being the amount that possibly could be dispersed
by plants and thereby affect the environment. The metal content in the roots is
not likely to be released into the environment, but will instead become a part of
the organic layer when the roots die. Metals may bind firmly to organic matter
(Schnoor, 1996). It is more difficult to predict what will happen with As.
The highest release of the metals was found in controls of Kristineberg
tailings (II), where also the lowest pH and highest sulphate concentrations were
found, which indicates a high weathering rate of sulphides. Arsenic was not
affected by the low pH in controls of Kristineberg tailings (Figs 8b, 11), but the
As levels were highest in the water from Kristineberg tailings.
In general, when comparing the same treatment in all different tailings, the
metal and As levels in drainage water from Kristineberg tailings (II) were
highest. This was also probably due to the high weathering rate and low
buffering capacity of Kristineberg tailings. One exception was the high Cu
content in water from Aitik tailing, due to the weathering of chalcopyrite
(Table 2), which contributed both to an elevated release of Cu to the drainage
water and a large Cu uptake and translocation to the shoot. Furthermore, in
plant treatments, the water from Laisvall and Garpenberg tailing had high Pb
content, likely due to the high Pb concentrations of those tailings.
Figure 11 shows that a great amount of the total possible release was found
in the shoots of the plants. The shoot contents of metals and As were generally
higher in E. angustifolium grown in the four tailings (Boliden, Garpenberg,
Aitik and Laisvall) used in paper IV than in the Kristineberg tailings used in
paper II. This is due to the greater biomass produced by the larger amount of
nutrients added in the paper IV experiments compared with those in paper II.
shoot content
283 µg
drainage water content
Cd (µg)
As (µg)
5287 µg
SS+E. a
SS+E. a
SS+E. a
SS+E. a
79926 µg
SS+E. a
SS+E. a
SS+E. a
SS+E. a
SS+E. a
SS+E. a
Zn (µg)
Pb (µg)
Cu (µg)
3811 µg
2085 µg
Figure 11. Amount of metals and As that could be dispersed by plants
(content in drainage water + shoot) from pots with different tailings and
treatments. Tailings: K=Kristineberg (II), B=Boliden, G=Garpenberg, A=Aitik,
L=Laisvall (IV). Treatments: C=controls, SS=sewage sludge, SS+E.a=sewage
sludge + E. angustifolium. Data from paper II were multiplied by 6.5 due to
differences in element accumulation time (see 2.7 for details).
High shoot contents of metal and As might be harmful to grazing animals,
especially since E. angustifolium is considered to be an important food source
for many animals in Northern Sweden (Warenberg, 1997). However, the shoot
metal and As concentrations of E. angustifolium has in general not been found
to be either toxic to animals or much higher than that of a ‘normal’plant (see
3.2.3; I). Furthermore, the major part of the shoots will probably not be
consumed by animals, but will instead wilt in the autumn before decomposing
and eventually becoming a part of the organic substrate. Studies have shown
that Cd, Cu and Zn remained in Salix leaves more than 100 days after
senescence (Greger et al., 2001). The mobility of the metals is dependent on
the pH of the covering water of the tailings, if it is acidic the metals are
mobilised, whereas if the pH is higher the metals will remain bound to the
organic matter (Schnoor, 1996). Less is known about the properties of arsenic
binding to organic matter.
When comparing the metals and As released that could affect the
environment (element content in drainage water + shoots) between E.
angustifolium, C. rostrata and P. australis (III), few differences were found.
No difference in the release of Cu, Fe and Pb was found between the plant
species. Eriophorum angustifolium had the highest release of Zn and Cd while
P. australis had the lowest. The release of As was lowest for E. angustifolium.
Hence, from the point of view of metals and As release that might affect the
environment all species seem similarly suitable for plant establishment on mine
3.3.5 Influences of different plant species on the total release of
metals and As from mine tailings rich in sulphides
The total release of metals and As (content in drainage water + shoots + roots)
might vary with different plant species. In paper III, differences in metals and
As release from sulphide-rich tailings were studied in pots with C. rostrata, E.
angustifolium and P. australis. The total amount of released metals and As is
dependent on the action of plants in the tailings. Plants may release the metals
and As from the tailings either by root uptake and translocation to the shoots,
or through subsequent release into the drainage water. The amount of released
metals and As by the plant, and where they will end up, may be due to the
plant’s ability to release elements from the tailings, its ability to accumulate
elements in tissue and may also depend on the biomass of the plant. Thus, a
plant with low biomass but with a great ability to accumulate metals and As
may take up more than a plant species with greater biomass but a lower
accumulation ability. Many plant species with hyperaccumulating properties
have low biomass (Reeves and Baker, 2000).
The differences found were that, even though P. australis had relatively
low shoot biomass and shoot concentration of Fe and As, it still gave the
highest total release of those elements caused by the high root and drainage
water concentrations (III). Carex rostrata and E. angustifolium had similar
total metal and As release but there was a tendency that C. rostrata had higher
element release in the drainage water and lower shoot concentrations than E.
angustifolium. Furthermore, C. rostrata and E. angustifolium had higher Cd
and Zn concentrations in the drainage water than P. australis.
The differences between P. australis and the other two species were
suggested to be due to differences in O2 release by the roots, thereby changing
the redox potential. The low redox potential found in pots with P. australis
may contribute to the formation of As(III) that is more mobile than As(V)
(Masscheleyn, 1991). In addition, As can bind to Fe oxides or hydroxides
formed in the upper part of the pots with higher redox potential but when they
are transported down with the drainage water to the reduced conditions, those
elements are released (Sracek, 2004). Phragmites australis is known to be able
to release large amounts of O2 by the roots (Armstrong, 1991; Brix, 1993),
which is the opposite of the results reported in paper II. The reason for the low
redox potential might be due to the experiment being performed in a
greenhouse where the plants were not exposed to wind. Differences in wind
velocity contribute to a great part of the oxidation of the rhizosphere of P.
australis, as discussed in paper III. Thus, in the field, the results might have
been different and the release of As and Fe might not have been so high as
found in paper III. The higher redox potential in pots with C. rostrata, and E.
angustifolium than in pots with P. australis might have increased the
weathering of sulphides, thereby causing the higher release of Cd and Zn (III).
Similar results have been found in other studies where Typha latifolia caused
higher Zn levels in the pore water of tailings than Glyceria fluitans (Wright and
Otte, 1999), possibly also due to the weathering of sulphide minerals since a
lower pH and an increased redox potential was found in tailings with T.
3.3.6 Metal and As release from mine tailings - long-term perspective?
There might be differences in metal and As release between the plant species
investigated (C. rostrata, E. angustifolium and P. australis), due to differences
in growth that were not seen in the experiments during a relatively short time
(13 months) (III). Greater root and shoot biomass might result in an increased
release of metals and As, and P. australis is known to be capable of reaching a
shoot height of 4 m (Cronk and Fennessy, 2001). Consequently, it is probable
that the shoot biomass is usually greater than the biomass found in paper III.
Then again, climate conditions also have a great impact. The plants of
Phragmites australis that had established naturally on mine tailings at the
Boliden mine site were found to have shoot heights of 1.5 m (Stoltz, pers.
observations). In addition, the root growth in the tailings was shallow,
approximately 0.4 m deep. Furthermore, as long as the conditions and nutrient
availability are sufficient for the plants, the roots will probably remain in the
amendments. In Danish wetland studies, roots of P. australis with high shoot
biomass had been observed as deep as 0.75 m below the soil surface, however,
most commonly the major part of the biomass was found in the upper 0.4 m of
the soil (Brix, 1998). Accordingly, even if a plant such as P. australis would
have a higher biomass than in paper III the plants will only affect a small part
of the volume of the tailings, since the thickness of the tailings within an
impoundment are around 4-11 m in Sweden (Carlsson, 2002). Thereby, the
effect of plants may not have severe adverse affects on the element release, as
also discussed by Jacob and Otte (2003).
If plant establishment decreases the metal release and increases the pH, as
has been shown for unweathered tailings with high content of sulphides, metals
and As, and low buffering capacity, phytostabilisation will be a suitable
remediation technique.
4 Conclusions
This work contributes with new knowledge about influences of plants growing
on water-covered mine tailings on element release, concluded by the following:
Plant establishment on water-covered mine tailings, both weathered and
unweathered, is possible. The shoot concentrations of metals and As in wetland
plants were in most cases found to be higher in roots than in shoots, thereby the
dispersion of these elements from the shoots should be smaller than if the
plants had been shoot accumulators.
The influence of E. angustifolium on metal and As release in the drainage
water varied with tailings. In tailings with very low levels of sulphides, or
higher levels of proton consuming minerals than sulphides, plants were in
general found to have few effects on the release of metals and As. The plant
species reduced metal levels in the drainage water and prevented a pH decrease
in tailings with high sulphide levels and low buffering capacity. However, in
tailings with lower levels of sulphides, the levels of metals in the drainage
water were increased, and E. angustifolium decreased the pH. No effect of E.
angustifolium was found on the release of As.
The different effects given by the plant species were suggested to be due to
the plants being able to cope with released elements up to a certain level. Plants
were able to cope with the released amount of metals and As from tailings with
relatively low sulphide levels and no pH increase was observed. In contrast, in
tailings with high sulphide levels, the released metal and As levels became
higher than the plant species could cope with, and to reduce the solubility of
the metals it would be necessary for the plant to induce some mechanism to
increase the pH. However, a plant mechanism to increase the pH was not found
and will need further investigations. Even though E. angustifolium was found
to be able to prevent a pH decrease in unweathered tailings, this species was
still unable to increase the pH of already weathered tailings.
Plant establishment on mine tailings has an aesthetic value, and few adverse
influences on the metal and As release in the drainage water were found. Also,
the release of elements from plant tissue is suggested to be limited. In sulphide
rich tailings with low buffering capacity plants increased the pH and reduced
the metal release, thus, phytostabilisation may be a successful way to remediate
such kind of mine tailings.
I would like to express my gratitude to:
♦ My supervisor, Associate Prof. Maria Greger, for support,
encouragement and giving me the opportunity to do a PhD in an
interesting subject.
♦ The ‘Plant metal group’: Åsa, Clara, Johanna, Yaodong and To mmy,
for interesting discussions (not only scientific), encouragement and
being such good friends.
♦ My co-supervisor Prof. Lena Kaustsky, for giving comments on my
♦ Former members of the ‘Plant metal group’: Lottie, Lasse B, Lasse E,
Martina, Lisa and Ann Helén for help and support. Extra special thanks
to Lisa and supervisor Maria for taking care of, and ending two of my
experiments when I became a mother earlier than expected.
♦ Patrik Dinnetz for statistical advice.
♦ The staff at the Department of Botany, all very nice and helpful.
♦ MiMi-people, especially Lars-Olof Höglund, for giving me the
opportunity to be part of the MiMi-project, and also the PhD students,
for interesting discussions and good meetings.
♦ Wiking Pettersson for help with the start-up of the field experiments
and collection of amendments.
♦ Dr. Marinus Otte for giving me the opportunity to visit the College
University of Dublin.
♦ Dr. Donna Jacob for taking care of me in Dublin and for demonstrating
her methods and showing me how to make a redox potential probe.
♦ The MISTRA financed project MiMi (Mitigation of the Environmental
Impact from Mining Waste), for financing my work.
♦ Boliden AB, for providing me with material and letting me work at
their mine sites.
♦ Mamma and Pappa, Fredrik and friends for being supportive and good
♦ Finally to the man in my life João, not only for all the help with
computers, proof-reading my thesis and putting up with me those last
stressful years, but also for being a good partner in life, and together
with our beautiful daughter Alice, making my days happy and
Armstrong, J., Armstrong, W. and Beckett, P. M. 1992. Phragmites australis: Venturi- and
humidity-induced pressure flows enhance rhizome aeration and rhizosphere oxidation.
New Phytol. 120: 107-207.
Baker, A. J. M. 1987. Metal tolerance. New Phytol. 106: 93-111
Baker, A. J. M., McGrath, S. P., Reeves, R. D. and Smith, J. A. C. 2000. Metal
hyperaccumulator plants: A review of the ecology and physiology of a biological
resource for phytoremediation of metal-polluted soils. In: Terry, N. and Bañuelos, G
(Eds), Phytoremediation of Contaminated Soil and Water. CRC Press, Boca Raton,
Baker, A. J. M. and Walker, P. L. 1990. Ecophysiology of metal uptake by tolerant plants. In:
Shaw, A. J. (Ed.). Heavy Metal Tolerance in Plants: Evolutionary Aspects. CRC Press,
Boca Raton, Florida, pp. 155-177.
Bergholm, J. and Steen, E. 1989. Vegetation establishment on a deposit of zinc mine wastes.
Environ. Pollut. 56: 127-144.
Bergström, J. 1997. Development of a geophysical method for investigating and monitoring the
integrity of sealing layers on mining waste deposits, final report. AFR-report 164.
Swedish Environmental Protection Agency.
Borgegård, S. O. and Rydin, H. 1989. Utilization of waste products and inorganic fertilizer in
the restoration of iron-mine tailings. J. Appl. Ecol. 26: 1083-1088.
Brieger, G., Wells, J. R. and Martin, M. H. 1992. Content in fly ash ecosystem. Water Air Soil
Pollut. 63: 87-103.
Brix, H. 1998. Denmark. In: Vymazal, J., Brix, H., Cooper, P. F., Green M. B. and Haberl, R.
(Eds.). Constructed Wetlands for Wastewater Treatment in Europe. Backhuys
Publishers, Leiden, The Netherlands, pp 123-152.
Brix, H. 1993. Macrophyte-mediated oxygen transfer in wetlands: Transport mechanisms and
rates. In: Moshiri, G. A. (Ed.), Constructed Wetland for Water Quality Improvement,
Boca Raton, London, Tokyo, pp. 391-398.
Burckhard, S. R., Schwab, A. P. and Banks, M. K. 1995. The effects of organic acids on the
leaching of heavy metals from mine tailings. J. Hazard. Mater. 41: 135-145.
Burd, G. I., Dixon, D. G. and Glick, B. R. 2000. Plant growth-promoting bacteria that decrease
heavy metal toxicity in plants. Can. J. Microbiol. 46: 237-245.
Carlsson, E. 2002. Sulphide-Rich Tailings Remediated by Soil Cover – Evaluation of cover
efficiency and tailings geochemistry, Kristineberg, northern Sweden. Luleå University
of Technology, Sweden, Doctoral thesis 2002:44, ISSN: 1402-1544.
Chen, R. L. and Barko, J. W., 1988. Effects of freshwater macrophytes on sediment chemistry.
J. Freshwater Ecol. 4: 279-289.
Cheong, T-W., Min, J-S and Kwon, K-S. 1998. Metal removal efficiencies of substrates for
treating acid mine drainage of the Dalsung mine, South Korea. J. Geochem. Explor. 64:
Clemens, S. 2001. Review: Molecular mechanisms of plant metal tolerance and homeostasis.
Planta 212: 475-486.
Clemensson-Lindell, A., Borgegård, S-O. and Persson, H. 1992. Reclamation of mine waste
and its effects on plant growth and root development – a literature review. Swedish
University of Agricultural Sciences. Report 47.
Coombes, D. M. 1997. The role of narrow-leaved cotton grass (Eriophorum angustifolium) in
the removal of copper in a sedge fen receiving acid mine drainage. Master of Sciences in
the School of Earth and Ocean Sciences, University of Victoria, Canada.
Coughtrey, P. J. and Martin, M. H., 1978. Cadmium uptake and distribution in tolerant and
nontolerant populations of Holcus lanatus grown in solution culture. Oikos 30: 555-560.
Cronk, J. K. and Fennessy, M. S. 2001. Wetland plants – Biology and Ecology. Lewis
publishers, Boca Raton, Florida.
Cunningham, S. D. and Berti, W. R. 1993. Remediation of contaminated soils with green
plants: An overview. Vitro Cell. Dev. Biol. 29: 207-212.
Drever, J. I. 1994. The effect of land plants on weathering rates of silicate minerals. Geochim.
Cosmochim. Acta. 58: 2325-2332.
Elander, P., Lindvall, M. and Håkansson, K. 1998. MiMi – Prevention and control of pollution
from mining waste products. MiMi Print, Sweden. ISBN 91-89350-02-2.
Ernst, W. H. O, Verkleij, J. A. C. and Schat, H. 1992. Metal tolerance in plants. Acta Bot.
Neerl. 41: 229-248.
Glick, B. R. 2003 Phytoremediation: synergistic use of plants and bacteria to clean up the
environment. Biotechnol. Adv. 21: 383-393.
Greger, M. 1999. Metal availability and bioconcentration in plants. In: Heavy Metal Stress in
Plants – from Molecules to Ecosystems. Prasad, M. N. V and Hagemeyer, J. (Eds.).
Springer-Verlag Berlin Heidelberg, Germany.
Greger, M., Brammer, E., Lindberg, S., Larsson, G. and Idestam-Almquist, J. 1991. Uptake
and physological effects of cadmium in sugar beet (Beta vulgaris) related to mineral
provision. J. Exp. Bot. 42: 729-737.
Greger, M., Ekvall, L., Österås, A-L., Perttu, K., Aronsson, P. and Pettersson, W. 1998. Mixed
waste products from the pulp and paper industry used as fertilizers in forest. Swedish
Environmental Protection Agency, AFR-report 238, pp. 7-12.
Greger, M., Landberg, T. and Berg, B. 2001. Salix clones with different properties to
accumulate heavy metals for production of biomass. Akademitryck Ab, Edsbruk,
Sweden, ISBN 91-631-1493-3.
Greipsson, S. and Crowder, A. A. 1992. Amelioration of copper and nickel toxicity by iron
plaque on roots of rice (Oryza sativa). Can. J. Bot. 70: 824-830.
Grimalt, J. O., Ferrer, M. and Macpherson, E. 1999. The mine tailing accident in Aznalcollar.
Sci. Total Environ. 242: 3-11.
Harmens, H., Koevoets, P. L. M., Verkleij, J. A. C. and Ernst, W. H. O. 1994. The role of low
molecular weight organic acids in the mechanism of increased zinc tolerance in Silene
vulgaris (Moench) Garcke. New Phytol. 126: 615-621.
Haynes, P. J. 1990. Active ion uptake and maintenance of cation-anion balance: A critical
examination of their role in regulation rhisosphere pH. Plant Soil 126: 247-264.
Hisinger, P., Barros, O. N. F., Benedetti, M. F., Noack, Y. and Callot, G. 2001. Plant-induced
weathering of a basaltic rock: Experimental evidence. Geochim. Cosmochim. Acta 65:
Hoffland, E., Findenegg, G. R. and Nelemans, J. A. 1989. Solubilization of rock phosphate by
rape. II. Local root exudation of organic acids as a response on P starvation. Plant Soil
113: 161-165.
Holmström, H. 2000. Geochemical processes in sulphidic mine tailings – field and laboratory
studies performed in northern Sweden at the Laver, Stekenjokk and Kristineberg minesites. Luleå University of Technology, Sweden, doctoral thesis 2000:03, ISSN: 14021544.
Holmström, H., Ljungberg, J., Ekström, M. and Öhlander, B. 1999. Secondary copper
enrichment in tailings at the Laver mine, northern Sweden. Environ. Geol. 38: 327-342.
Jacob, D.L and Otte, M.L. 2003. Conflicting processes in the wetland plant rhizosphere: metal
retention or mobilisation? Water Air Soil Pollut. 3: 91-104.
Jacob, D. L. and Otte, M. L. 2004a. Influence of Typha latifolia and fertilization on metal
mobility in two different Pb-Zn mine tailings types. Sci. Total Environ. 333: 9-24.
Jacob, D. L. and Otte, M. 2004b. Long-term effects of submergence and wetland vegetation om
metals in a 90-year old abandoned Pb-Zn mine tailings pond. Environ. Pollut. 130: 337345.
Jones, D. L. 1998. Organic acids in the rhizosphere – a critical review. Plant Soil 205: 25-44.
Jones, D. L. and Darrah, P. R. 1994. Role of root derived organic acids in the mobilization of
nutrients from the rhizosphere. Plant Soil 166: 247-257.
Jung, M. C. 2001. Heavy metal contamination of soils and waters in and around the Imcheon
Au-Ag mine, Korea. Appl. Geochem. 16: 1369-1375.
Kelly, E. F., Chadwick, O. A. and Hiliniski, T. E. 1998. The effect of plants on mineral
weathering. Biogeochem. 42: 21-53.
Laan, P. Smolders, D. Blom, C. W. P. M. and Armstrong, W. 1989. The relative roles of
internal aeration, radial oxygen losses, iron exclusion and nutrient balances in floodtolerance of Rumex species. Acta Bot. Neerl. 38: 131-145.
Landberg, T. and Greger, M. 1996. Differences in uptake and tolerance to heavy metal in Salix
from unpolluted and polluted areas. Appl. Geochem. 11: 175-180.
Lowson, R. T. 1982. Aqueous oxidation of pyrite by molecular oxygen. Chem. Rev. 82: 461497.
Ma, J. F., Ryan, P. R. and Delhaize, E. 2001. Aluminium tolerance in plants and the
complexing role of organic acids. Trends Plant Sci. 6: 273-278.
Marschner, H. 1995. Mineral Nutrition of Higher Plants. Academic Press, London.
Marschner, H. and Römheld, V. 1983. In vivo measurement of root-induced pH changes at the
soil-root interface: Effect of plant species and nitrogen source. Z. Planzenphysiol. Bd.
111: 241-251.
Masscheleyn, P. H, Delaune, R. D and Patrick Jr, W. H. 1991. Effect of redox potential and pH
on arsenic speciation and solubility in a contaminated soil. Environ. Sci. Technol. 25:
Matthews, D. J, Moran, B. M., McCabe, P. F. and Otte, M. L. 2004. Zinc tolerance, uptake,
accumulation and distribution in plants and protoplasts of five European populations of
the wetland grass Glyceria fluitans. Aquat. Bot. 80: 39-52.
McCabe, O. M. and Otte, M. L. 2000. Revegetation of metal mine tailings under wetland
conditions. In: Brandt J. E., Galevotic, J. R., Kost, L., Trouart, J. (eds) Proceedings of
the 14th Annual National Meeting – Vision 2000: An Environmental Commitment,
American Society for Surface Mining and Reclamation, Austin, Texas
McDowell, L. R. 1992. Minerals in animal and human nutrition. Academic Press, San Diego,
Milton, A. and Johnson, M. 1999. Arsenic in the food chains of a revegetated metalliferous
mine tailings pond. Chemosphere 39: 765-779.
MiMi (2001): Mitigation of the Environmental Impact from Mining
Waste (MiMi), Programme Plan for the period 2001-2003, ISBN 91-89350-12-X.
NRC (National Research Council), 1980. Mineral Tolerance of Domestic Animals. National
Academy of Science, Washington, D.C., USA. ISBN 0-309-03022-6.
Nixdorf, B., Fyson, A. and Krumbeck, H. 2001. Review: plant life in extremely acidic waters.
Environ. Exp. Bot. 46: 203-211.
Nyavor, K., Egiebor, N. O. and Fedorak, P. M. 1996. Suppression of microbial pyrite oxidation
by fatty acid amine treatment. Sci. Total Environ. 182: 75-83.
Nye, P. H. 1981. Changes of pH across the rhizosphere induced by roots. Plant Soil 61: 7-26.
Ochs, M. 1996. Influence of humified and non-humified natural organic compounds on mineral
dissolution. Chem. Geol. 132: 119-124.
Otte, M. L. and Ernst, W. H. O. 1994. Arsenic in vegetation of wetlands. In: Arsenic in the
Environment, Part 1: Cycling and Characterization. John Wiley and Sons Inc., New
Otte, M. L., Rozema, J., Koster, L., Haarsma, M. S. and Broekman, R. A. 1989. Iron plaque on
root of Aster tripolium L.: interation with zinc uptake. New Pytol. 111: 309-317.
Pais, I. and Jones Jr., J. B. 2000. The Handbook of Trace Elements. CRC Press, Boca Raton,
Raskin, I. and Ensley, B. D. (Eds.). 2000. Phytoremediation of toxic metals: using plants to
clean up the environment. John Wiley & Sons, New York.
Reeves, R. D. and Baker, A. J. M. 2000. Metal-accumulation plants In Phytoremediation of
Toxic Metals using Plants to Clean Up the Environment. Eds Rasking, I. and Ensley,
B. D. John Wiley and Sons Inc., New York, pp 193-230.
Salmon, S. U. 2003. Geochemical Modelling of Acid Mine Drainage in Mill Tailings. KTH
Land and Water Resources Engineering, Stockholm, Sweden, Doctoral Thesis, ISBN
Salt, D. E., Blaylock, M., Kumar N. P. B. A., Dushenkov, V., Ensley, B. D., Chet, I. and
Raskin, I. 1995. Phytoremediation: A novel strategy for the removal of toxic metals
from the environment using plants. Bio-technol. 13: 468-474.
Sracek, O., Bhattacharya, P., Jacks, G., Gustafsson, J-P. and von Brömssen, M. 2004. Behavior
of arsenic and geochemical modelling of arsenic enrichment in aqueous environments.
Appl. Geochem. 19:169-180.
Schnoor, J. L. 1996. Modeling trace metals. In: Environmental modelling – Fate and transport
of pollutants in water, air, and soil, pp 381-451. John Wiley & Sons, Inc. U.S.
Sneller, F. E. C., Van Heerwaarden, L. M., Kraaijeveld-Smit, F. J. L., Bookum, W. M.,
Koevoets, P. L. M., Schat, H. and Verkleij, J. A. C. 1999. Toxicity of arsenate in Silene
vulgaris, accumulation and degradation of arsenate-induced phytochelatins. New Phytol.
144: 223-232.
Sobolewski, A. 1999. A review of processes responsible for metal removal in wetlands treating
contaminated mine drainage. Int. J. Phytoremediat. 1: 19-51.
Strömberg, B. and Eriksson, N. 1996. Vittringsegenskaper hos anrikningssand: En analys av
fuktkammarförsök med anrikningssand från Aitik. Avdelning för vattenvårdsteknik,
institutionen för anläggning och miljö, The Royal Institute of Technology, TRITA-AMI
REPORT 3022. ISSN 1400-1306. In Swedish.
Thompson, K., Grime, J. P. and Mason, G. 1977 Seed germination in response to diurnal
fluctuation of temperature. Nature 267: 147-149.
Tordoff, G. M., Baker, A. J. M. and Willis A. J. 2000. Current approaches to the revegetation
and reclamation of metalliferous mine wastes. Chemosphere 41: 219-228.
Thurman, C. A. and Rankin, J. L. 1982. The role of organic acids in zinc tolerance in
Deschampsia caespitosa. New Phytol. 91: 629-635.
Villegas, J. and Fortin, J. A. 2001. Phosphorus solubilization and pH changes as a result of the
interactions between soil bacteria and arbuscular mycorrhizal fungi on a medium
containing NH4+ as nitrogen source. Can. J. Bot. 79: 865-870.
Warenberg, K., Danell, Ö., Gaare, E. and Nieminen, M. 1997. Flora i Renbetesland. Werner
Söderström Osakeyhtiö-WSOY, Finland. In Swedish.
Wasay, S. A., Barrington, S. and Tokunaga, S. 2001. Organic acid for the in situ remediation of
soils polluted by heavy metals: soil flushing in columns. Water Air Soil Pollut. 127:
Whitbread-Abrutat, P. H. 1997. The potential of some soil amendments to improve tree growth
on metalliferous mine wastes. Plant Soil 192: 199-217.
Wilkinson, F., Beckett, P. J. and St-Germain, P. 1999. Establishment of wetland plants on
flooded mine tailings. Mine, Water & Environment, IMWA Congress, Sevilla, Spain.
Wright, D. and Otte, M. L. 1999. Wetland plant effects on the biogeochemistry of metals
beyond the rhizosphere. Biology and environment: Proceedings of the Royal Irish
Academy, Vol 99B, No. 1: 3-10.
Wu, L. 1990. Colonization and establishment of plants in contaminated sites. In: Heavy metal
tolerance in plants: Evolutionary aspects, Ed. Shaw, A. J. CRC Press Inc. Boca Raton,
Florida, pp. 269-284.
Ye, Z. H., Wong, J. W. C, Wong M. H., Lan, C. Y. and Baker, A. J. M. 1999. Lime and pig
manure as ameliorants for revegetating lead/zinc mine tailings: a greenhouse study.
Biores. Technol. 69: 35-43.
Ye, Z. H., Yang Z. Y., Chan, G. Y. S. and Wong, M. H. 2001. Growth response of Sesbania
rostrata and S. Cannabina to sludge-amended lead/zinc mine tailings A greenhouse
study. Environ. Int. 26: 449-455.
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