Phytostabilization of mine tailings covered with fly ash and sewage sludge Clara Neuschütz

Phytostabilization of mine tailings covered with fly ash and sewage sludge Clara Neuschütz
Phytostabilization of mine tailings
covered with fly ash and sewage
sludge
Clara Neuschütz
©Clara Neuschütz, Stockholm 2009
ISBN 978-91-7155-807-7, pp 1-53
Front cover: Phalaris arundinacea growing in sewage sludge
on top of a fly ash sealing layer. Boliden, Sweden. Photo:
Clara Neuschütz.
Printed in Sweden by Universitetsservice, US-AB, Stockholm
2009
Distributor: Department of Botany, Stockholm University
Abstract
Establishing plant communities is essential for the restoration of contaminated land. As potential cover materials, fly ash and sewage sludge can prevent formation of acid mine drainage from sulfidic mine waste. The aim of
the thesis was to i) screen for plants that can be established in, and prevent
leakage of metals and nutrients from sludge on top of ash and tailings, and ii)
investigate root growth into sealing layers of ash and sludge. Analyses were
performed under laboratory, greenhouse and field conditions using selected
plant species to examine the release of Cd, Cu, Zn, N, and P from the materials. Plant physiological responses and interactions with fly ash were also
investigated.
The data show that plants can decrease metal and nutrient leakage from
the materials, and lower the elemental levels in the leachate, but with varying efficiencies among plant species. Plants capable of taking up both nitrate
and ammonium were more efficient in preventing N leakage compared with
those taking up primarily ammonium. Fast growing plants could raise the pH
in acidic sludge leachate, but the initial pH decrease and N leakage was not
counteracted by plants. Germination in fresh sludge was problematic, but
enhanced by aeration of the sludge. In general, the accumulation of metals in
plant shoots was low, especially if ash was located below the sludge. Fresh
ash was phytotoxic (e.g., high alkalinity, salinity and metal levels) and induced the activity of stress-related enzymes in shoots. In sealing layers of
aged and cured ash, roots could grow if the penetration resistance was low,
or into the surface of stronger layers if the surface had become pulverized.
The roots caused dissolution of calcium-rich minerals, possibly by exudation
of saccharides. Addition of sludge to an ash layer increased root growth,
likely due to decreased bulk density and pH, and nutrient addition. In conclusion, with selected plant species and a properly constructed cover, metal
and nutrient leaching from the materials and root growth into the sealing
layer can be restricted.
List of papers
This thesis is based on the following papers, which will be referred to by
their Roman numerals:
I.
Neuschütz, C. and Greger, M. Ability of various plant species to prevent N, P, and metal leakage from sewage sludge.
Accepted for publication by International Journal of Phytoremediation.
II.
Neuschütz, C. and Greger, M. Stabilization of mine tailings
using fly ash and sewage sludge planted with Phalaris arundinacea L. Submitted to Water, Air, and Soil Pollution.
III.
Neuschütz, C., Stoltz, E. and Greger, M. 2006. Root penetration of sealing layers made of fly ash and sewage sludge. Journal of Environmental Quality 35:1260-1268.
IV.
Neuschütz, C., Boström, D. and Greger, M. Root growth
into sealing layers of cured fly ash. Manuscript.
My contributions to the papers were as follows: I was responsible for the
writing of all papers, with help from the co-authors. Together with my coauthors I planned all experiments, with the exception of one of the root penetration studies in Paper III. The field and laboratory work was performed by
me, except application of material in the field and some elemental and mineral analysis, which are mentioned in the respective papers.
Paper I and III are reproduced with kind permission from the publishers.
Contents
1 Introduction ..............................................................................................7
1.1 Phytostabilization.................................................................................................. 7
1.2 Mine waste and acid mine drainage (AMD) .................................................... 7
1.3 Dry cover treatment of mine tailings ............................................................... 9
1.4 Sewage sludge .................................................................................................... 10
1.5 Fly ash................................................................................................................... 10
1.6 Ecophysiological aspects of plant establishment......................................... 12
2 Aim............................................................................................................14
3 Comments on materials and methods ..............................................15
3.1 Field sites ............................................................................................................. 15
3.2 Plant species used .............................................................................................. 16
3.3 Waste products ................................................................................................... 18
3.4 Analysis of nutrients and metals..................................................................... 19
3.5 Determination of penetration resistance....................................................... 20
4 Results and Discussion .........................................................................21
4.1 Plant growth in sewage sludge and effect on the leachate....................... 21
4.1.1 Establishment of plants in sewage sludge ........................................... 21
4.1.2 Metal uptake in plants .............................................................................. 22
4.1.3 Leakage of nutrients ................................................................................. 25
4.1.4 Leakage of metals ..................................................................................... 29
4.2 Root penetration of sealing layers .................................................................. 31
4.2.1 Plant growth in fly ash ............................................................................. 31
4.2.2 Penetration resistance of fly ash layers................................................ 33
4.2.3 Plant loosening effects on cured fly ash............................................... 34
4.3 Long term function and choice of plants ....................................................... 36
5 Conclusions .............................................................................................39
6 Future work .............................................................................................41
Sammanfattning på svenska ..................................................................42
Acknowledgements ...................................................................................44
References ..................................................................................................46
6
1 Introduction
1.1 Phytostabilization
Generally the use of plants in the restoration of contaminated sites is termed
phytoremediation, a technique that evolved during the last decades of the
20th century (Salt et al., 1995). When an area is only slightly contaminated,
plants with a strong ability to take up metals from soils can be used in a
process termed phytoextraction. However, when dealing with more heavily
polluted sites, like sites for the disposal of mine tailings, plants that do not
transport the metals to the shoots, but instead bind them in the root or the
rhizosphere, are preferred. This approach is termed phytostabilization
(Wong, 2003). The plants may also have beneficial effects on the treatment
by causing physical stabilization, which prevents erosion, and by preventing
rainwater from percolating down to underlying contaminated soil or mine
tailings (Tordoff et al., 2000).
Introduction of plants directly on mine tailings has repeatedly been attempted, but has usually failed. This is due to the fact that such impoundments offer a harsh environment with high levels of heavy metals, low levels
of macronutrients and poor substrate structure (Smith and Bradshaw, 1979;
Clemensson-Lindell et al., 1992; Bell, 2001). Therefore, ameliorative materials like sewage sludge have been added in many cases, resulting in increased water holding capacity, reduced phytotoxicity and provision of nutrients over a long period of time (Hearing et al., 2000; Tordoff et al., 2000).
The role of and function of plants in covers consisting of deep layers of fly
ash and sewage sludge, however, remains to be examined.
1.2 Mine waste and acid mine drainage (AMD)
Extraction of metals by mining results in large volumes of wastes that have
to be taken care of in order to avoid contamination of the environment. The
problem arises primarily after extraction of non ferrous base metals like Cu,
Pb and Zn, since these are found in ores with high sulfide content. After the
metals have been extracted, approximately 95% of the rock is left as finely
7
grained sand called mine tailings, containing high levels of metal sulfides
(Table 1), among which pyrite (FeS2) is most abundant.
When mine waste containing pyrite comes into contact with oxygen and
water, dissolved iron and sulfuric acid are produced through several complex
reactions, of which one of the most important is the overall reaction (1)
(Rimstidt and Vaughan, 2003):
FeS2 + H2O + 3.5 O2 → Fe2+ + 2 H+ + 2 SO42-
(1)
The Fe(II) ions formed may generate more acid through further oxidation,
and turn into Fe(III), which may act as an oxidant on pyrite. The weathering
process results in strongly acidic drainage water, commonly referred to as
acid mine drainage (AMD), in which heavy metals can dissolve easily. Oxidation of other sulfides may also generate free metal ions as shown in reaction 2; however, this reaction does not generate any protons (Holmström,
2000):
MeS + 2 O2 → Me2+ + SO42-
(2)
Table 1. Levels of elements in mine tailings, sewage sludge, fly ashes and common
soil, and tolerable levels (of metals, As and B) for plants. Elements not analyzed are
marked with “-“.
Element
(g kg-1
DW)
Ca
Fe
K
Mg
Na
P
As
B
Cd
Cu
Pb
S
Zn
1
Mine
tailings1
0.18
0.02
1.0
0.5
144
9.0
Sewage
sludge2
MSW fly
ashes 3
28
49
4.4
3.4
3.5
27
0.005
0.06
0.001
0.4
0.03
9.0
0.6
74-130
12-44
22-62
11-19
15-57
4.8-9.6
0.04-0.3
0.05-0.5
0.6-3.2
5.3-26
11-45
9-70
Biofuel fly Background Tolerable
ashes4 levels in soil5 levels for
plants6
230
14
15
33
90
27
45
7.4
22
15
10
1.1
0.01
0.004
0.05
0.3
0.005
0.1
0.01
0.0002
0.005
0.1
0.02
0.1
0.2
0.02
0.1
11
0.4
1.7
0.07
0.3
Unoxidized mine tailings (73 samples) from Kristineberg, Sweden (Holmström et al., 2001)
Average values from 48 Swedish sewage treatment plants in the year 2000 (Eriksson, 2001)
3
Range of content in fly ash from incineration of municipal solid waste (MSW) (Wiles, 1996)
4
Biofuel fly ashes from 12 Swedish power stations (Westermark and Gromulski, 1996)
5
Topsoils (0-20 cm below soil surface) throughout Sweden (25 samples) (Eriksson, 2001)
6
Proposed plant-tolerable levels of trace elements in soil (Pais and Jones, 2000)
2
8
Microorganisms present in the mine waste will affect the processes and increase or decrease the rate of metal mobilization (as reviewed by Ledin and
Pedersen, 1996).
Each year approximately 20 million tons of reactive sulfidic mine tailings
that must be disposed of are generated in Sweden (Fröberg and Höglund,
2004). About half of that amount can be used as filling mass in terminated
mines, but the rest has to be stored in large impoundments (Södermark,
1986), for which natural lakes are often used (Fig. 1). If these mine tailings
are left uncovered, the finely grained sand may be spread by the wind and
contaminate the surrounding environment. Furthermore, the upper layer (0.51.5 m) will with time become oxidized, releasing sulfide-bound metals to
surrounding waters (Holmström et al., 1999). In Sweden, where the bedrock
has a low buffering capacity, the acidic drainage water from mine tailings
may be a source of heavy metal leakage to various waters (Notter, 1993).
Therefore, much effort has been invested in the development of techniques
to limit these environmental problems.
Figure 1. Mine tailings impoundment without cover, at Lake Gillervattnet, Boliden.
The total area of the lake is ~250 ha (Larsson et al., 2005) (Photo: Clara Neuschütz).
1.3 Dry cover treatment of mine tailings
One way to restrict the formation of AMD is to construct a dry cover that
prevents weathering of the mine tailings either by acting as a barrier against
oxygen and/or precipitating water, or as a consumer of oxygen (Elander et
al., 1998). Beneficial is also if it increases the ground water table. A sufficient dry cover can be based on one material or a combination of different
substrates. A multi-layer cover should consist of at least one compact sealing
layer, and one more porous layer, which protects the sealing layer from erosion and temperature fluctuations and forms a suitable substrate for vegetation (Fig. 2). Traditionally, moraine has commonly been used to cover mine
9
tailings, but since the nutrient values are low and the extraction of moraine is
expensive and detrimental to the landscape, new materials are under evaluation (Clemensson-Lindell et al., 1992). Waste materials from other industrial
activities are attractive, since they are available at low cost. Two such products are sewage sludge from waste water treatment plants and fly ash from
thermal power stations.
1.4 Sewage sludge
Sewage sludge is the solid material remaining from the treatment of waste
water (Fytili and Zabaniotou, 2008). The treatment commonly includes a
primary (chemical and physical treatment), secondary (biological treatment),
and tertiary step (removal of nutrients). In Sweden each year, approximately
210 000 tons of sewage sludge are produced (Statistics Sweden, 2008). The
composition is highly variable and levels of trace elements like Cd, Zn, Cu,
and Pb (Table 1) may be high. Sewage sludge with low levels of pollutants
is, however, considered to be a valuable plant fertilizer, due to its good water-holding capacity and high nutrient content (Tordoff et al., 2000). In restored industrial areas, like old mining sites, sewage sludge has frequently
been used as an ameliorating cover material (Sopper, 1993; Hearing et al.,
2000); in Sweden this application has garnered attention as agricultural use
declines. Another term for sewage sludge is “biosolids”, which is often used
for sludge low in pathogens, and which is considered suitable as soil
amendment and fertilizer (WEF, 2009).
Apart from being a suitable substrate for plant establishment, sewage
sludge may function well as a component of sealing layers (Mácsik et al.,
2003), since the sludge in a compacted form can become almost impermeable to air and water (Scandiaconsult Sverige AB, 2001). However, sewage
sludge alone does not have enough strength and durability to function as a
sealing layer, and should therefore be used in combination with fly ash.
Concerns that have been raised about the use of sewage sludge in dry covers
are related to leakage of eutrophicating nutrients and toxic metals, emission
of greenhouse gases (e.g., nitrous oxide and methane) and dispersal of
pathogens. This thesis has focused on plant effects on nutrient and metal
release.
1.5 Fly ash
Fly ash is produced during combustion of biofuel, coal, oil or municipal
solid waste (MSW), and consists of particles that are captured from the
smoke by filters (Wiles, 1996). Each year in Sweden, 435 000 tons of fly ash
are produced and must be disposed of (Ribbing, 2007). The composition of
10
fly ash varies greatly depending on combusted material. Generally, fly ash
contains high levels of Si, Ca, Al, S and a number of heavy metals (Table 1,
Steenari et al., 1999). Properties that make fly ash suitable as a material in
sealing layers include high alkalinity (Adriano et al., 1980), possibly representing a capacity to neutralize acidic drainage water, as well as the solidifying properties enabling ash to harden after being mixed with water (Steenari
et al., 1999).
The hardening, or stabilizing, process is due to high levels of lime (CaO)
and calcium sulfate (CaSO4), which in contact with water form strong crystalline structures, in reactions similar to those that characterize the hardening
of concrete (Steenari et al., 1999). The reactions include (from Steenari and
Lindqvist, 1997):
ƒ
ƒ
Carbonatization, e.g., Ca(OH)2 (portlandite) + CO2 → CaCO3 (calcite) + H2O
Formation of
• Hydroxides, e.g., CaO + H2O → Ca(OH)2
• Gypsum, CaSO4 + 2H2O → CaSO4·2H2O
• Ettringite, e.g., Ca3Al2O6 + 3 CaSO4·2H2O + 26H2O →
Ca6Al2(SO4)3(OH)12 ·26 H2O
• Hydrated silicate and aluminum-silicate phases
The initial crystallization is rapid, whereas further stabilization reactions
may continue for months, depending on the composition of the fly ash
(Steenari et al., 1999). The strength of the hardened fly ashes seems to be
correlated to water ratio (Xie and Xi, 2001) and to be favored by compaction
of the material (Steenari et al., 1999) and high levels of calcium and available silica (Sivapullaiah et al., 1998).
Issues related to the use of fly ash in dry covers include its ability to
harden and form a layer with low hydraulic permeability, as well as the risk
of toxic leakage. Some ashes may alone form layers with low enough permeability (Tham and Andreas, 2008), while others need addition of other
materials to decrease the permeability, e.g., sewage sludge (Mácsik et al.,
2003), lime (Scheetz and Earle, 1998; Prashanth et al., 2001) or bentonite
(Mollamahmutoglu and Yilmaz, 2001). In particular, municipal solid waste
(MSW) ashes may contain high levels of Cd, Pb and Zn (Table 1). MSW
ashes are considered to be hazardous waste in Sweden, making utilization in
landfill covers controversial. However, in fly ash applied in the field weathering reactions may immobilize metals and prevent them from leaching
(Wiles, 1996). Fly ash has even been used to decrease metal leaching from
contaminated soils, by absorbing the metals on oxide and hydroxide surfaces
as well as by increasing the pH (Stouraiti et al., 2002).
In Sweden, the possibilities of recycled ash have been studied in the context of the “Environmentally-friendly use of non-coal ashes” project, initiated in 2002 by the Swedish Thermal Engineering Research Institute (Vär-
11
meforsk). This has resulted in an increased use of ashes for landfills; construction of roads and parking places; forestry and mine tailing coverings.
Approximately 50% of all ashes produced in Sweden were utilized in 2003;
however, only a smaller fraction as cover material on mine tailings (Ribbing,
2007).
1.6 Ecophysiological aspects of plant
establishment
For successful establishment of vegetation in a dry cover treatment, the
plants must tolerate the conditions in the protective cover, but avoid growing
into and penetrating the sealing layer (Fig. 2).
A
B
Protective cover
Sealing layer
Mine tailings
No release of
nutrients or
metals
Leakage of
nutrients
and metals and
formation
of AMD
Figure 2. Possible roles of vegetation in a dry cover treatment of mine tailings, using
sewage sludge and fly ash. The optimal situation is shown in picture A), where the
plants prevent leakage of nutrients and metals, and do not grow into the sealing
layer. In picture B) plants do not prevent nutrient and metal leakage from the sludge,
but rather penetrate the sealing layer, leading to oxidation of the mine tailings and
formation of acid mine drainage (AMD).
Since sewage sludge and fly ash do not fully resemble any natural soils, suitable plant species have to be selected after examination of their responses to
the materials. Natural soils with a high content of moisture and organic matter (such as bogs) resemble sewage sludge in some aspects. However, since
they are usually formed in cold climates with slow nutrient turnover (Brady
12
and Weil, 2002), plants growing in these soils are adapted to low nutrient
levels and would not be optimal for establishment in sewage sludge. The
underlying fly ash layer may in some ways resemble soils formed by deposition of volcanic ash (Andisols), with crystalline structures and high pH
(Brady and Weil, 2002). These soils have, however, been exposed to soilforming processes for several thousands of years, and are much less reactive
than newly applied fly ash. Hypothetically, plants that would avoid growing
into an ash sealing layer are those sensitive toward alkaline and calcium-rich
soils (so called calcifuges). Nevertheless, these plants are generally adapted
to soils with low salinity and nutrient content (Larcher, 2003), and may not
tolerate the conditions in sewage sludge. Appropriate plants would, instead,
be those with a high nutrient uptake and a weak ability to grow into compacted soil.
An ability to utilize nitrogen both in the form of ammonium (NH4) and nitrate (NO3) should be beneficial, since both forms will occur in sewage
sludge after application to land. Most plant species have an optimal growth
with a mixture of ammonium and nitrate (Marschner 1995), however, the
preferences may vary between plants adapted to different climates and soil
conditions. Plants adapted to cold climates, where the nutrient turnover rate
is slow and soils often have a low pH, generally have a higher ability to take
up ammonium than nitrate, and plants growing in Ca-rich and more alkaline
environments have a higher preference for nitrate (Gigon and Rorison 1972).
Furthermore, plants occurring early in the succession (such as many ruderate
species) often have a greater ability to adapt to variations in nitrogen species
(Min et al. 2000), while woody plants occurring late in the succession usually are dependent on organically bound nitrogen and ammonium (Nordin et
al. 2001).
In re-vegetation of mine waste covered with sewage sludge, it has been
common to introduce various grass species with tolerance toward metals and
salinity (Bendfeldt et al., 2001; Evanylo et al., 2005; Santibáñez et al., 2008).
It may be assumed that grasses have shallower root systems than woody
plants. However, rooting depth has been more strongly related to climatic
variables than to vegetation types (Schenk and Jackson, 2002). Therefore,
further studies on physiological traits rendering plants suitable for growth in
sewage sludge and fly ash should be performed, to optimize the use of plants
in the restoration of mining sites.
13
2 Aim
The aim was to increase our knowledge of the interaction between plants and
the waste product sewage sludge and fly ash used in treatment of sulfidic
mine tailings. I examined physiological responses of plants growing in the
materials and mechanisms behind plant-induced impacts on the materials.
The goal was, further, to improve our understanding of the extent to which
plants can be used in phytostabilization of mine waste, and how sewage
sludge and fly ashes can be used in an environmentally safe way.
The project was divided in two parts:
1.
The establishment of plants in sewage sludge and their effect on
leakage of nutrients and metals. The specific aim was to reveal:
a. plant species that are capable of establishment in sewage sludge in
a temperate climate
b. the effect of various plant species on leakage of nutrients and metals from sewage sludge, fly ash, and mine tailings, and the mechanisms involved
c. the extent to which plants growing in a cover of sewage sludge and
fly ash accumulate metals in their shoots, posing a hazard to grazing animals
2.
Root penetration of sealing layers containing fly ash. The specific aim
was to determine:
a. how plants respond physiologically toward fly ashes with various
properties
b. the properties and mechanical resistance necessary in an ash sealing layer to prevent penetration of plant roots
c. whether plants can affect the strength of a sealing layer of fly ash,
for instance by exudation of root substances
Experiments were performed within a time frame of days to several months;
root growth was observed in the field for up to eight years. The results are
presented in the four papers attached. Included in the thesis are also data
from preliminary studies concerning plant establishment in sewage sludge,
uptake and distribution of metals in plants, and concentrations of N and P in
sewage sludge pore water.
14
3 Comments on materials and methods
3.1 Field sites
The field experiments (Paper IV) were
performed at the mine tailings impoundments at Gillervattnet, Boliden (64°52’N,
20°22’E) (Fig. 3). Observations were also
performed at Garpenberg (60°19'N,
Boliden
16°09'E), in test plots for a project initiated
by the municipality of Hedemora together
with Boliden AB and Högskolan Dalarna.
Various cover treatments have been conGarpenberg
structed at both these sites, using waste
↑
products from wastewater treatment plants,
N
thermal power stations and the wood indus300 km
try, in order to evaluate the possibilities of
utilizing these materials.
At the tailings impoundment in Boliden,
Figure 3. Field work has been
performed at Boliden and Garsealing layers consisting of either 100%
penberg (Paper IV). Marked
biofuel fly ash (from Munksund and Heareas are regions in Sweden
densbyn), 70% fly ash and 30% sewage
with sulfidic bedrock.
sludge (based on volume), or 100% sewage
sludge were constructed, with depths of 0.5 to 1.0 m. Sludge has been applied at depths of 0.25 to 0.5 m on top of these sealing layers (Fig. 4). The
test plots have been built gradually and vegetation introduced during the
years 2003 to 2006. These plots have been used for investigations of plant
establishment, uptake of metals in shoots, pore water chemistry in the sewage sludge, and root penetration of sealing layers.
At the tailings impoundment at Västra sandmagasinet, Garpenberg, test
plots were constructed in 1998, with waste products from the local paper
board industry (Stora Enso Fors). A 30 cm deep layer of fly ash was applied,
covered with a 45 cm thick layer of paper pulp sludge mixed with mine tailings. Introduction of various tree species was performed in 1999 and 2000,
whereas the root growth was studied in autumn 2007.
15
Figure 4. Vegetated test plots in sewage sludge on top of a sealing layer of fly ash at
the mine tailings impoundments in Boliden. (Photo: Clara Neuschütz).
3.2 Plant species used
Plant species suitable for establishment in dry covers of sewage sludge and
fly ash on mine tailings, should be tolerant to the materials and able to prevent leakage of nutrients and metals without accumulating metals in their
shoots. Moreover, the plant roots should have a weak ability to penetrate
compacted fly ash layers. The plants also need to tolerate a colder climate,
since most mining areas in Sweden are situated in the north or middle part of
the country (Fig. 3). The mean temperature of Västerbotten province, where
Boliden is situated, ranges from -6°C to -15°C in January and from 12 to
15°C in July; annual precipitation is 500 to 700 mm (Vedin, 2007). The
Dalarna province, where Garpenberg is located, has a mean temperature in
January of -5.5°C to -12°C and in July of 10°C to 16°C, and an annual precipitation of 600 to 1000 mm (Vedin, 2005). Different types of plants, such
as herbs, trees and grasses with varying anatomical and physiological capacities were included in the analyses (Table 2, Fig. 5). Certain plants are of
particular interest since they may establish through natural succession, while
others can easily be manually dispersed by seeds. In addition, crops that can
be used for biofuel may be of future interest and were included in the project. To find plant species that are adapted to the conditions in these areas, an
inventory of naturally dispersed plant species was performed at a sewage
sludge disposal site next to the mine tailings in Boliden, as well as at a site
next to the mine tailings in Garpenberg, where a thin layer of untreated sewage sludge has continued to spread over 30 years. Out of more than 70 plant
species found growing naturally in both areas, a smaller number were chosen
for later studies (Table 2, Fig. 5).
16
Table 2. Plant species used in the experiments presented in the thesis. Their lifehistory, life-form, and minimum rooting depth are given (USDA, NRCS 2009).
Latin name
Common name
Life-history and lifeform
Agrostis capillaris L.*
Betula pendula Roth.*
Cannabis sativa L.
Epilobium angustifolium
L.*
Phalaris arundinacea L.
Phaseolus vulgaris L.
Picea abies (L.) H.
Karst.*
Pinus sylvestris L.*
Pisum sativum L.
Poa pratensis L.*
Common bent
Silver birch
Industrial hemp
Fireweed
Perennial grass
30
Perennial deciduous tree 60
Annual herb
Perennial herb
25
Salix myrsinifolia
Salisb.*
Salix viminalis L.
Tussilago farfara L.*
Min. root
depth (cm)
Reed canary grass Perennial grass
35
Dwarf bean
Annual herb
Norway spruce
Perennial coniferous tree 70
Scots pine
Garden pea
Kentucky bluegrass
Dark-leaved willow
Basket willow
Coldsfoot
Perennial coniferous tree 50
Annual herb
Perennial grass
25
Perennial deciduous
shrub
Perennial deciduous tree
Perennial herb
*Plant species found growing naturally at the mining sites
a
b
c
d
Figure 5. Plant species used for the leakage studies (Papers I and II): a) Epilobium
angustifolium, b) Phalaris arundinacea, c) Pinus sylvestris, and d) Salix viminalis.
(Photo: Clara Neuschütz).
Phalaris arundinacea, Salix viminalis, and Cannabis sativa were selected to
represent energy crops in Sweden (Venendaal et al., 1997). Cultivation could
be of economic interest, and these plant species are known to have a high
nutrient uptake. However, S. viminalis has a relatively low tolerance toward
frost (Tahvanainen and Rytkönen, 1999), and may therefore be difficult to
cultivate in the northernmost parts of Sweden.
In the toxicity test of fly ash extracts in Paper III, Pisum sativum was
used, although this is not a plant intended for establishment at treated mining
areas. Advantages of using this plant in short-term toxicity tests are the fast
growth rate of primary roots and the nutrient-rich seed, making addition of
17
nutrients to the solution unnecessary. A toxicity test performed in the same
way but with 19-day-old P. sylvestris seedlings in dilution series of two of
the fly ashes showed that these plants were more sensitive toward the ashes
than P. sativum. Similarly, in the study of physiological stress response toward diluted fly ash extracts (Paper III), Phaseolus vulgaris was used because it has previously been used in standardized biological test systems
(Van Assche and Clijsters, 1990).
When comparing different plant species, one has to consider variations in
growth pattern and life-history. An annual herb may, for instance, increase
its biomass faster during the first growing season than a perennial tree, while
the tree will have higher biomass within a couple of years. In many cases,
plants of the same age can be used; if the biomass differs, this can be accounted for as a species difference. However, when roots are for instance
washed for the collection of root exudates, a large variation in root biomass
can be problematic if the same amount of washing liquid is used. The diffusion of elements from the roots could in that case be affected. To avoid such
errors, I have in some experiments used plants with similar biomass. This
has been performed either by varying the time of pre-cultivation, or by varying the number of plants in each container. In the field study, seeds, cuttings
and one to several-year-old plantlets were used to study the ability to establish and grow into the sealing layer.
3.3 Waste products
Fly ashes were received from seven different thermal power stations using
varying combustion techniques and types of fuel (Table 3). In a circulating
fluidized bed (CFB) furnace, finely grained sand is used as bed material to
homogenize the heat distribution, thus reducing the temperature requirements (800°C instead of conventional 1450°C) (Scheetz and Earle, 1998).
The amount of Si in ashes from CFB combustors is usually higher than from
traditional grating furnaces, due to residual bed material particles. This can
have an impact on hardening capacity, since higher puzzolanic ability has
been related to higher levels of available Si (Sivapullaiah et al., 1998).
The sewage sludge that was used in greenhouse and field studies originated from the Henriksdal sewage treatment plant, Stockholm Vatten,
Stockholm, which treats a large fraction of all sewage produced by households in Stockholm. Iron sulfate is used as a precipitation agent at this treatment plant; the sludge is stabilized unaerobically and dewatered to a dry
matter content of about 30%.
18
Table 3. Origin and type of fuel and furnace of the fly ashes used in the thesis.
Origin
Fuel
Hedensbyn, Skellefteå 80% biofuel, 20% peat
Kraft
Hässelby
Biofuel
Furnace
CFB1
Paper
III, IV
Pulverized fuel
III
Högdalen
Iggesund Paperboard
Munksund, SCA
Packaging
Stora Enso Fors Mill
Umeå Energi
CFB1
III
Grate-fired boiler IV
CFB1
II, III, IV
1
Wood construction waste
Bark
91% biofuel, 6% recycled
paper, 3% oil
Biofuel
Municipal solid waste
CFB1
II, IV
Grate-fired boiler III, IV
CFB = Circulating Fluidized Bed
3.4 Analysis of nutrients and metals
The analysis of nutrients in leachates (Papers I and II) was performed spectrophotometrically, at 640 nm for NH4-N, 220 nm for NO3-N (Clesceri and
Greenberg, 1995), and at 880 nm for PO4-P (Murphy and Riley, 1962). Before analysis the samples were filtered (<0.45µm), resulting in a determination of dissolved species but not total content. Interference due to organic
matter can be a problem during spectrophotometric analysis of nitrate. The
high nitrate concentrations, however, made dilution possible, so that the
problem was avoided.
The reason for analyzing the metals Cd, Cu, and Zn in leachate (Papers I
and II) was their high abundance in waste materials (Table 1) and their toxicity to water organisms (Fjällborg et al., 2005; Madoni and Romeo, 2006),
which may affect organisms downstream from the restored mining site.
Analysis was performed using an atomic absorption spectrophotometer
(AAS, Varian SpectraAA-100, Springvale, Australia), with flame technique
for Zn and furnace (GTA-97) for Cd and Cu. Three standards were added to
each sample to correct for interaction of the sample matrix. Upon analysis of
plant metal content, we included certified reference material (CRM) from P.
arundinacea (NJV 94-4) and Salix (NJV 94-3), originating from the Swedish
University of Agricultural Sciences, Uppsala, Sweden, in order to validate
the digestion procedure and to ensure high-quality metal analysis.
The amount of metals in the sewage sludge (Paper II) was analyzed after
digestion in 7M HNO3 (for 30 minutes at 120°C) (according to Swedish
Standard SS 028150). This method does not dissolve the metal fraction
bound to silicate minerals, however, extracting close to the total amount of
metals from rich organic substrates (Sastre et al., 2002).
19
3.5 Determination of penetration resistance
The resistance that a root is encountering in a soil can be measured using a
penetrometer. This device records either the pressure required to press a steel
probe (usually with a tip semiangle of 15° or 30°) to a certain depth or at a
certain rate, or the depth of penetration under a constant load (Bennie, 1991).
The result will not completely capture the resistance that the root is exposed
to, since no apparatus can imitate all features of a root, such as exudation of
lubricating mucilage and radial growth of cells. The penetrometer used in
Papers II and III was constructed at the Department of Soil Science, Swedish
University of Agricultural Sciences, Uppsala. The cone had a semiangle of
15°, a diameter of 4.1 mm, and was mounted on a 3 mm wide and 30 mm
long piston on an apparatus originally constructed for measurement of soil
aggregate strength (Fig. 6).
Force
Figure 6. Schematic picture of the penetrometer used for determining penetration resistance in cured fly ash (Paper II,
IV).
By determining the maximal force that was needed to drive the cone 10 mm
down into the material, the penetration resistance was calculated. One problem with using a laboratory penetrometer is that only smaller samples can be
measured. Smaller aggregates will crack earlier than larger aggregates
(Misra et al., 1986), causing a lower resistance than would be the case in a
homogenous layer in the field. Therefore, the penetration resistance of fly
ash from the field (Paper IV) may be underestimated. In the greenhouse
study, however, the penetration resistance was measured in fly ash layers in
planting pots maintained under the same conditions as those experienced by
roots encountering the fly ash. For the experiments in Paper III, we used the
bulk density of the fly ash samples as a measure of mechanical impedance,
since the penetrometer had not been constructed at that time. The penetration
resistance depends partly on the bulk density (Bennie, 1991; Vaz et al.,
2001), which has been used as a measure of soil strength in several studies of
plant responses to soil compaction (Nasr and Selles, 1995; Stirzaker et al.,
1996; Arvidsson, 1999). Other important factors are water content of the soil
and porosity (Bennie, 1991), which was not analyzed in the study in Paper
III, but was investigated in Paper IV. The occurrence of biopores has been
found to greatly enhance the possibilities of root growth in soils with high
bulk density (Stirzaker et al., 1996).
20
4 Results and Discussion
4.1 Plant growth in sewage sludge and effect on
the leachate
4.1.1 Establishment of plants in sewage sludge
The high levels of nutrients in sewage sludge make it a suitable substrate for
plants. However, even though it contains all essential plant nutrients, the
proportions are not ideal for most plants. Often a surplus of P and/or a deficiency of K is the issue (Bramryd, 2002). Moreover, the establishment of
plants in sewage sludge can be a problem due to high salinity (Abad et al.,
2001), high levels of metals, and ammonium (Fjällborg and Dave, 2003).
This can be ameliorated if the sludge is mixed with soil or peat (GarciaGomez et al., 2002). A decrease in pH, due to reactions such as nitrification,
decomposition of organic S, or oxidation of Fe sulfides (Merrington et al.,
2003), may cause an increase in availability of toxic metals, but can be
avoided by addition of a liming substrate (Sajwan et al., 2003).
Germination and the establishment of seeds or seedlings in fresh sewage
sludge was often problematic, both in greenhouse (Paper I, II) and field experiments (Paper IV). A hard sludge surface caused by dry weather could
further obstruct germination. A germination test (unpublished) showed that
germination and growth during the first three weeks in sewage sludge was
considerably lower than in planting soil (Fig. 6). The plant species used
were: Agrostis capillaris. Betula pendula, Epilobium angustifolium, Phalaris
arundinacea, Poa pratensis, and Pinus sylvestris. Even cuttings of Salix
myrsinifolia that were transferred to pots with pure sewage sludge had difficulties in establishing as compared to those transferred to planting soil (Fig.
7). However, once the plants have passed this stage and survived the first
months they display robust growth in the sludge.
To ensure successful establishment of plants in the greenhouse experiments, the sludge needed to be dispersed and exposed to air for a couple of
weeks. The plants were also in some cases pre-cultivated in planting soil
until roots had developed (Paper II). Thereafter they were transferred to
21
140
120
Soil (control)
100
80
60
40
Sludge
*
*
*
*
S.
m
ra
te
ns
e
P.
p
a
ng
us
t if
ol
iu
m
P.
ar
un
di
na
ce
a
P.
sy
lv
es
tri
s
en
du
l
E.
a
B.
p
ap
i ll
A.
c
yr
si
ni
fo
lia
*
20
0
ar
is
Shoot DW in % of control
sewage sludge. In field experiments, we observed that a layer of bark between the sealing layer of fly ash and the protective sludge cover increased
the rate of survival one month after transferral of plantlets (of Betula pubescens, Picea abies and Pinus sylvestris), compared to plantlets established in
only sludge. On average, 74% of all plants survived when bark was added
compared with 19% in only sludge (unpublished data). Germination of P.
arundinacea was also more efficient when bark was present, while no effect
was seen in S. myrsinifolia and S. caprea. It is possible that the bark improved the conditions by increasing the oxygen levels in the otherwise compacted sludge, as well as protecting against toxic effects from the fly ash
below. With time composting and degradation processes will occur in the
sludge, making it more tolerable for plant germination (Fuentes et al., 2006),
for instance by decreasing the ammonium levels by nitrification (Smith and
Tibbett, 2004).
Figure 7. Shoot biomass (dry weight in % of control) of various plant species three
weeks after seeding in planting soil (control) or sewage sludge. Salix myrsinifolia
was introduced as cuttings and harvested after 4 weeks (n=5, +SE) (unpublished
data). Significant differences (p<0.05) between sludge treatment and control are
marked with a star.
Thus, to rapidly establish plants in fresh sewage sludge, the sludge should be
aerated, or a material such as bark can be added. As a dry sludge surface
should be avoided, application and establishment of plants should preferably
be performed during spring or autumn.
4.1.2 Metal uptake in plants
From other studies, the plant uptake of metals from sewage sludge has been
found moderate and not exceeding levels that are toxic for plants (Miller et
al., 1995; Pichtel and Anderson, 1997; Evanylo et al., 2005) or animals
(Stuczynski et al., 2007). However, addition of a liming material has been
22
recommended (Brown et al., 2003; Sajwan et al., 2003) to prevent pH decrease in the sludge that otherwise can increase metal phytoavailability (Antoniadis et al., 2008).
In this thesis, analysis of metal uptake was performed in plants that had
been growing for 2.5 months in sewage sludge alone, or on top of fly ash and
mine tailings (Papers I and II) (Table 4). Furthermore, metal content was
determined in shoots of five different plant species collected in the field after
growth in sludge on top of a fly ash sealing layer for up to 3 years. The shoot
concentrations of Cd were low in all cases. Concerning Cu and Zn, the concentrations reached the maximum tolerable levels for animal feed (NRC,
2005) in some of the plants (Table 4). These levels are, however, recommendations for domestic animals ingesting no other food (NRC, 2005).
Higher levels should be tolerable for wild animals that can move freely and
graze over large areas.
Table 4. Average levels of Cd, Cu, and Zn in shoot tissue of various plant species
used in the field (unpublished) and greenhouse experiments (Papers I and II). The
recommended maximum tolerable levels in feed for domestic animals (mg kg-1 air
dried forage) (NRC, 2005) are also presented.
Plant species
Cannabis sativa
Epilobium angustifolium
Phalaris arundinacea
Poa annua
Pinus sylvestris
Salix myrsinifolia
Salix viminalis
Tolerable levels for
animals
Field (F) or
greenhouse
(G) study
F
G
F+G
F
G
F
F+G
Cd
Cu
Zn
0.3-0.4
24
7-18
11-20
17
6-8
13
11-22
111
94-190
135-342
115
36-78
1090
140-290
10
15-500
300-1000
0.2-0.4
0.1-0.2
0.04-0.1
Similarly high shoot Zn levels have been found in Salix species growing in
mine tailings at Boliden, explained by the high capacity of the plants to
translocate metals to the shoots (Stoltz and Greger, 2002). In one of the
greenhouse experiments the uptake of Cd, Cu and Zn was analyzed in P.
arundinacea (Paper II). If looking at the distribution of Cd, Cu and Zn
among roots, shoots, and sludge after harvest, it turned out that various metals were distributed differently (Fig. 8) (unpublished data). While shoot uptake of Cd and Cu was restricted, transport of Zn to the shoots was higher.
This study also showed that growth in sewage sludge did not significantly
increase metal uptake as compared with growth in planting soil (containing
80% peat and 20% clay). Furthermore, the plants were not able to change the
23
metal concentrations in the sludge during the experiment, and the accumulation factor (metal concentration in the plant tissue divided by metal concentration in the sludge) was below 1 in all cases. This means that the plants did
not accumulate a higher concentration of metals in their tissue than was
found in the surrounding substrate (not shown). The highest accumulation
factors were observed for Zn (0.52 in shoots, and 0.77 in roots). Plants used
in phytostabilization should preferably exhibit weak accumulation of metals
in shoots, to avoid transferring toxic metals into the food chain.
100%
80%
Shoot
60%
Root
40%
Sludge
20%
0%
Cd
Cu
Zn
Figure 8. Distribution of metals in sewage sludge, roots and shoots of P. arundinacea after 11 weeks of growth (n=7) (unpublished data).
In the field, we found that Cu and Zn concentrations in shoots of P. arundinacea that had been growing in sludge for 0.5 to 3 years did not vary significantly among the years (not shown). Thus, either the plants maintained a
constant level of uptake, or the availability of these metals stayed constant
over the years.
The presence of fly ash below the sewage sludge prevented Cu and Zn
uptake in roots of P. arundinacea, and in shoots when mine tailings were
also included (Paper II). It is likely that the alkalinity of the fly ash decreased
the availability of metals, since many metals are less soluble at higher pH (in
the region close to neutral) (Villar and Garcia, 2002; Basta et al., 2005).
Furthermore, fly ashes contribute Mn and Fe oxides, minerals that have been
suggested to be responsible for decreased uptake of metals in plants grown
in fly ash-stabilized sewage sludge (Su and Wong, 2003).
The presence of mine tailings below the sludge had a minor effect on
metal uptake in plants (Paper II). Only Cu was somewhat higher in shoots
when the tailings were present. Other studies have evaluated the effect on
metal uptake of adding sewage sludge to mine tailings, with varying results.
In some cases, metal uptake in plants has decreased when sludge is added
(Theodoratos et al., 2000; Ye et al., 2001), as compared to the absence of
organic amendment; in other cases, uptake has increased (Rate et al., 2004;
Santibáñez et al., 2008). The result has also been found to vary depending on
the weathering status of the tailings, with a preventive effect of sludge on
24
plant metal uptake from oxidized tailings, but an increasing effect on uptake
from non-oxidized tailings (Stjernman Forsberg, 2008).
In conclusion, the introduction of plants in a protective cover of sewage
sludge on top of fly ash does not expose grazing animals to toxic levels of
Cd, Cu or Zn, provided that certain metal-accumulating plant species are
avoided.
4.1.3 Leakage of nutrients
PO4-P leakage (µg week
-1
)
Application of sewage sludge at mine disposal sites poses a risk of nutrient
leakage to the surroundings, especially of nitrogen; phosphorus has been
found firmly bound in the sludge (Paper I, Stehouwer et al., 2006). Plant
establishment can decrease leakage (Santibáñez et al., 2007). However, since
the availability of nutrients from the sludge does not necessarily match the
nutrient needs of the plants (Bramryd, 2002), superfluous elements can be
subjected to leaching.
Data obtained from the greenhouse experiments demonstrate that plant
uptake of water is one of the most important factors in leakage prevention,
since it will decrease the total amount of leachate formed (Papers I and II).
Water uptake can also cause aeration of the sludge, rendering phosphorus
more firmly bound to FeOOH (Gomez et al., 1999). In the leakage study
presented in Paper I, the total leakage of phosphate (Fig. 9) and nitrate, but
not ammonium, correlated negatively with the water uptake of plants.
100
80
60
40
R2 = 0.5127
20
0
0
0.05
0.1
0.15
Water uptake in plants (L w eek -1)
0.2
Figure 9. Total leakage of phosphate-P in
relation to water uptake per week of four
plant species growing
in containers with
sludge for 10 weeks.
(Paper I).
Hydrological conditions affected by plants are, thus, of great importance for
the stabilizing effect, and are schematically shown in Fig. 10. Fast-growing
plants with a high transpiration rate should theoretically be the most favorable plants; however, it was found that the solution is not that simple (Paper
I). Both Phalaris arundinacea and Salix viminalis are fast-growing energy
crops, with high water uptake and biomass production (Venendaal et al.,
1997). The concentrations of nutrients in the leachate varied, however,
among the species, with S. viminalis causing higher concentrations of nitrate
25
and ammonium, but lower levels of phosphate (at some weeks) compared
with P. arundinacea. The weed Epilobium angustifolium also had a high
growth rate and was almost as efficient as P. arundinacea in preventing
leakage of N and P. Such differences between plant species in the effect on
leachate may result from higher uptake of a certain element. Alternately,
plants may alter the conditions in the sludge, rendering elements less mobile
(e.g., by altering pH or redox status).
Air
Root
Sewage
H2O sludge
particle
NO3H+
NH4+
H+
Men+
Fe(II)
Water
phase
FeOOH-P
PO43-
Figure 10. Processes that may occur at
the interface of plant roots and sewage
sludge. Water uptake by the plant increases the diffusion of air, and can
thereby influence the equilibrium of nutrients and metals.
Plants taking up both ammonium and nitrate (E. angustifolium and P. arundinacea) were more efficient in preventing nitrogen leakage than those
mainly taking up ammonium (S. viminalis and Pinus sylvestris) (Paper I).
Most plants are able to take up N in several forms (mainly as NH4+ or NO3-,
but also in organic forms such as amino acids), and are usually benefited by
a mixture of species yielding a balance between negatively and positively
charged molecules (Marschner, 1995). Assimilation of N from NH4+ is favorable for plants, since that requires less energy than using other forms, but
NO3- is often the form that is most available (Marschner, 1995). Ideal plants
for establishment in sewage sludge are those with a high and adaptive uptake
of nitrogen.
While S. viminalis exhibited a weaker ability to prevent nitrate leaching
relative to P. arundinacea, the effect was reversed with regard to phosphate
(Paper I). In a supplementary experiment (unpublished data), the amount of
nutrients attached to the roots of these plants was examined, after growth in
filter pockets between sewage sludge for 48 hours. The roots were then
washed in deionized water, which was filtered (<0.45µm) and analyzed for
nitrate and phosphate, performed using an ion chromatograph (Dionex
26
ED50); and ammonium by spectrophotometry. The results showed that while
the amount of attached nitrate was higher on the roots of P. arundinacea,
roots of S. viminalis had higher amounts of attached phosphate (Table 5). It
is possible that plants with a high uptake of certain nutrients also have a
great ability to attach these nutrients to the roots, and thereby prevent their
leaching.
Table 5. Levels of nutrients in the root-washing medium; number of root tips per g
DW biomass; root biomass of Salix viminalis and Phalaris arundinacea (n=6, ± SE
(unpublished data). Letters indicate significant differences (p<0.05) between the
plants.
NH4-N (mg g-1)
NO3-N (mg g-1)
PO4-P (mg g-1)
Number of root tips (g-1)
Root biomass (g DW)
S. viminalis
2.2 ±0.4a
4.3 ±0.5a
0.30 ±0.08b
710 ±110 a
0.16 ±0.04a
P. arundinacea
3.1 ±1.3a
7.8 ±0.5b
0.03 ±0.02a
140 ±10 b
0.14 ±0.05a
The architecture of the root system is also likely to have an effect on the
nutrient uptake. Counting the number of root tips showed that S. viminalis
had significantly more root tips per root DW than P. arundinacea (Table 4).
This may explain the higher capacity for phosphate uptake in S. viminalis,
since many nutrients are taken up in the vicinity of the root tips (Marschner,
1995), where the suberinized layers in endodermis and exodermis have not
yet developed. Nitrate, on the other hand, is taken up throughout the entire
root of plants (Guo et al., 2001). This could explain the robust ability of P.
arundinacea to reduce nitrate leakage, compared to S. viminalis (Paper I),
despite a lower number of root tips per unit biomass (Table 4). Grasses have,
furthermore, been found to have higher N:P ratios than other plant types
(Eckstein and Karlsson, 1997). Therefore, grasses should be suitable for
filtering N rather than P, while S. viminalis could be more useful if the goal
is to reduce the release of P.
Results obtained from greenhouse studies in small containers, or in hydroponic solutions, will quite likely differ from the results obtained during
complex field conditions, as has been found by e.g., Stoltz and Greger
(2002). Preliminary studies were performed to evaluate the effect of vegetation on pore water in sewage sludge in the field (unpublished data), as an
indication of what was available for leaching. Pore water was collected by
vacuum suction into soil moisture samplers (MacroRhizon, Eijkelkamp,
Netherland), 9 cm long and 4.5 mm wide, composed of a porous inert material. The soil moisture samplers were placed in sludge planted with either 1)
P. arundinacea, 2) a grass mixture (Svalöf Weibull road slope mixture), 3)
weeds from natural succession (mainly E. angustifolium and S. myrsinifolia),
or 4) no vegetation, at the mine tailings impoundment at Gillervattnet, Boliden (Fig. 4). We also included sludge that had been applied annually (0.5 to
27
600
Spring
400
Autumn
a
a
ab
200
ab
ab
b
0
PO4-P (mg L-1)
NH4-N (mg L-1)
3 years before the study), and which had been planted with P. arundinacea,
to determine whether the duration of sludge exposure to field conditions had
an effect. Samples were taken in spring and in autumn and analyzed for pH
and levels of ammonium, nitrate and phosphate.
Establishment of vegetation did not cause any significant effects on the
composition of the pore water (not shown). Instead, the season at which
sampling was performed, and the age of the sludge influenced the pH and
nutrient concentrations (Fig. 11).
2
1
0.5
ab
b
2003 2004 2005 2006
pH
NO3-N (mg L-1)
abab
ab
ab
ab b
ab b
ab
2003 2004 2005 2006
a
ab
abab
0
2003 2004 2005 2006
3000
2500
2000
1500
1000
500
0
a
1.5
10
8
6
4
2
0
a
b b
ab
ab
b
b
b
2003 2004 2005 2006
Year of sludge application
Figure 11. pH and concentrations of ammonium, nitrate and phosphate in pore water
collected during the spring and autumn (31 May and 3 Oct) in 2006, from sewage
sludge that had been applied annually from 2003 to 2006. The sludge was planted
with Phalaris arundinacea (n=3, +SE) (unpublished data). Bars with different letters
(a-b) differ significantly (p<0.05) from each other.
While ammonium and phosphate levels were highest in newly applied
sludge, and decreased over the course of the season, there were indications
that nitrate concentration increased both over the course of the season and
with the age of the sludge. The pH decreased with time, indicating that a
nitrification process was occurring, transforming ammonium into nitrate, and
releasing protons (Merrington et al., 2003).
Nitrate levels were high in the pore water of sludge obtained from the
field (Fig. 11) and in leachate collected in the greenhouse (Paper I), far exceeding the water quality threshold value (NO3-N > 50 mg L-1) of the European Commission nitrate directive (91/676/EEC) (CEC, 1991b). The concentrations of phosphate were, on the other hand, low in all experiments,
remaining below the water quality threshold of the urban wastewater treatment directive (91/271/EEC): 2 mg L-1 (CEC, 1991a).
28
Neither field nor greenhouse experiments tell us how much will actually
leach from a complete system with sludge, fly ash and tailings. The materials
below the sludge will affect what will leach; therefore leachate should also
be collected at the outlet of the mine waste disposal site. Indications of how
combinations of the materials affected the leachate were, however, obtained
in the leachate study presented in Paper II. The presence of fly ash and mine
tailings had an impact on the leakage of nutrients from the sludge. The presence of fly ash, for instance, decreased the concentrations of ammonium,
nitrate and phosphate, at least when plants were also present (Paper II). In a
study of waste water sludge stabilization, fly ash demonstrated strong nitrate- and phosphate-adsorbing capabilities (Topaç et al., 2008). Mine waste
has, moreover, been found to bind nutrients from nutrient-rich materials and
to prevent leaching (Santibáñez et al., 2007; Wei et al., 2008), even though
no such effects were observed in our study (Paper II).
Although other factors than plant establishment seem to have a larger impact on leaching of N and P in the field, the greenhouse studies indicate that
plants can influence the leakage of nutrients. Fast-growing plants with the
ability to use both ammonium and nitrate seem to be the most suitable for
establishment, in order to prevent early leaching.
4.1.4 Leakage of metals
Sewage sludge, fly ash and mine tailings contain high levels of many metals
(Table 1). However, it is generally not the total metal content in a medium
that predicts the availability and subsequent risk of leaching, but rather factors such as pH and the levels of organic matter and Al, Fe, and Mn oxides
(Basta et al., 2005). Plants can greatly affect the conditions in the growth
substrate (Gregory, 2006), and may therefore influence metal leaching.
When grown in sewage sludge, with no underlying layers of fly ash or
mine tailings, different plant species (Fig. 5) displayed varying ability to
affect metal leaching (Paper I). Phalaris arundinacea was most efficient in
preventing leakage; Pinus sylvestris was least efficient. Even though P.
arundinacea exhibited lower uptake of Cd and Zn, and similar water uptake
rate compared with Salix viminalis, the former plant was more efficient in
preventing leakage of metals (Paper I). In the study of attachment of nutrients to roots (unpublished data, described in chapter 4.1.3 Leakage of nutrients), also the metal (Cd, Cu and Zn) content in the root-washing medium
was analyzed. From these data we found that P. arundinacea had significantly higher amounts of Cd and Zn attached to the roots (per dry weight)
than what S. viminalis had, while there was no difference concerning Cu. As
previously discussed regarding N and P, it is possible that an ability to attach
metals to the roots also results in a high capacity to prevent leakage of the
elements. To further study absorption of metals, we analyzed root cation
29
exchange capacity (root CEC) of the plants. This is a measure of the number
of negatively charged sites in the pectin in cell walls, acting as exchange
sites for positive ions from the soil solution (Haynes, 1980). A higher root
CEC has been suggested to result in higher uptake of cations in certain plants
(Han et al., 1998). However, the relationship is complex; root CEC may vary
with plant age, nutrient content (Heintze, 1961; Bakker and Nys, 1999), and
metals (Greger and Landberg, 2008) in the growth medium. In Paper I, positive correlations were found between root CEC and total amounts of leached
metals, but not with the plant uptake of metals. Theoretically, a plant with a
high root CEC and therefore a robust ability to liberate metals from the substrate could also increase the leakage of metals, if the uptake of metals in the
plant does not correspond to the amount mobilized from the sludge.
Combining sewage sludge, fly ash and mine tailings will make leaching
behavior more complex. Several studies have shown that fly ashes can prevent metal leaching from mine waste through physical and chemical mechanisms involving coating of pyrite grains with metal precipitates (PérezLópez et al., 2007) and alkalinization, as well as by adsorption of metals to
Al and Fe (hydro)oxides (Kumpiene et al., 2007). In Paper II, the leaching of
metals was studied in mine tailings covered with a thin and porous layer of
fly ash, covered with sludge, with and without plants. The presence of plants
yielded high levels of metals in the leachate; no preventive effect of fly ash
was observed. The porous layer of fly ash did not prevent root growth into
the mine tailings, which could start weathering after increased access of
oxygen due to water uptake by plants. Instead, the fly ash seemed to increase
plant growth, resulting in greater water uptake by plants and increased
weathering of the mine tailings than if no fly ash was present.
The effect of sewage sludge on metal leaching from mine tailings varies
among different studies and seems to depend on site-specific conditions.
Preventive effects are sometimes found. These have been related to the prevention of oxygen diffusion to mine tailings (Peppas et al., 2000), binding of
metals (Holtzclaw et al., 1978), or stimulation of sulfate-reducing bacteria,
producing pyrite on which metals can precipitate (Gibert et al., 2004). However, addition of sludge can also increase the mobilization of metals, by providing ligands in form of organic acids (Ribet et al., 1995), or humic and
fulvic acids, which may become mobilized from FeOOH with increasing pH
(Weng et al., 2007). The high nutrient content of sewage sludge can, furthermore, stimulate iron-oxidizing bacteria under aerobic conditions and
thereby increase pyrite oxidation (Cravotta, 1998).
In Paper II, the effect of placing either sludge or fly ash in direct contact
with the mine tailings was investigated over a time period of three months.
During this time, there was no difference in metal release between these two
treatments. However, as compared to uncovered mine tailings, both treatments significantly decreased leakage. Plant (P. arundinacea) growth decreased the amount of leachate, but did not affect the concentrations of met30
als in the leachate (Paper II). This implies that leakage may occur during the
winter season following diminished plant growth and water uptake.
In summary, different plants vary in their ability to prevent metal leakage
from sewage sludge (Paper I). This variation was mainly related to plant
uptake of water. If a sealing layer is used that cannot prevent root growth,
fast-growing plants may increase the access of oxygen to mine tailings and
release of metals (Paper II). Constructing a sealing layer that is resistant to
growing roots will prevent weathering of the mine tailings.
4.2 Root penetration of sealing layers
4.2.1 Plant growth in fly ash
Fly ash in a concentrated form generally harms plants, due to high levels of
salts (Giordano et al., 1983), boron (Wong et al., 1998; Gupta et al., 2002),
metals (Tripathi et al., 2004), or high alkalinity (Paper III, Gupta et al.,
2002). However, fly ash also contains most essential elements for plant
growth, except nitrogen, and can have positive effects on plants after becoming stabilized. Stabilized fly ash has been recommended for use in forestry,
especially in acidic soils (Naylor and Schmidt, 1986), where it can increase
available P, Ca, Mg, K and B, as well as decrease toxic effects of Al and Mn
by raising the pH (Demeyer et al., 2001).
During stabilization, the fly ash becomes hydrated and carbonated, processes that turn reactive CaO and Ca(OH)2 into CaCO3, decreasing the alkalinity. We have been able to show that aging of hardened fly ash for ten
months clearly made it less toxic toward plants, as compared with fly ash
that was hardened for one week (Paper III). Plants could not grow into fresh
ash, even in porous layers, while aged ash prevented root growth only when
hardened to a high density.
It is difficult to determine which factor has the most adverse effect on root
growth, in part due to variability among plant species. In the greenhouse test
of root penetration, presented in Paper IV, Pinus sylvestris had a higher
shoot biomass after growth on top of the metal-rich MSW fly ash, compared
with growth on an alkaline biofuel fly ash; the opposite was true for Epilobium angustifolium and Salix myrsinifolia. This indicates that P. sylvestris
has a high sensitivity toward alkalinity, while the other two species are more
sensitive toward high metal levels. The fourth plant tested, Phalaris arundinacea, exhibited high growth in all treatments, but also displayed morphological signs of stress in all ashes, such as stunted roots with enlarged cortex,
deformed cell walls and deformed root hairs (Fig. 13, Paper IV). Reduced
root elongation is a sign of stress due to mechanical impedance (Bennie,
1991); however, since roots were stunted even in the ash layers with low
31
penetration resistance (Paper IV), the effect is more likely to be due to
chemical parameters, such as alkalinity, or high levels of metals or salts.
a
b
Cortex
ca 100 µm
Cortex
100 µm
d
c
Cortex
ca 350 µm
100 µm
Figure 13. Root hairs and cross-sectioned roots of Phalaris arundinacea after
growth either in a sand/clay mixture (a, b) or in biofuel fly ash (c, d). Note the disturbed growth of cell walls and root hairs (c), and increased cortex diameter (d). The
roots were examined using light microscopy (a, c), and ESEM (environmental scanning electron microscopy (b, d) (Paper IV). (Photo: Clara Neuschütz).
Most likely, the combination of stress factors caused a so-called stressinduced morphogenic response. This has been described to include a) inhibition of cell elongation, b) localized stimulation of cell division and c) alteration in cell differentiation status (Potters et al., 2007). The theory is that
many stressors at sub-lethal levels induce a similar growth pattern – where
the tap root is inhibited and lateral roots are stimulated. This is a strategy to
avoid unfavorable soil, due to either chemical or mechanical stress, or a deficiency of nutrients. Important signals in this response are ROS (reactive
oxidative species) and phytohormones, particularly auxin (Potters et al.,
2007). By measuring the activity of stress-induced enzymes, such as peroxidases, stress can be analyzed in plants before growth is affected (Van Assche
and Clijsters, 1990). In a comparison of the effects of weak extracts from a
32
metal-rich MSW fly ash and a biofuel fly ash on plants (P. vulgaris), the
MSW ash gave rise to much higher activity of certain stress-related enzymes
in the shoots (Paper III). Thus, as alkalinity decreases (in this case by dilution of the ash extract), a high content of metals may obstruct root growth. In
the long term, however, the levels of available metals will decrease, leaving
mechanical resistance as the most important factor in preventing root
growth.
4.2.2 Penetration resistance of fly ash layers
Since the sealing layer aims to prevent diffusion of oxygen down to the tailings, it is of great importance that roots are not penetrating this layer. This
became obvious in the two experiments presented in Paper II, using either a
compacted or a porous layer of fly ash between the mine tailings and the
covering sewage sludge. The porous fly ash could not prevent roots from
growing into the mine tailings, where they caused desiccation, weathering of
the mine tailings and metal release (Paper II). Non-stabilized fly ash can
prevent root growth, by acting toxic due to high pH, metal or salt content, as
discussed in the previous chapter. However, as reactivity decreases with
time, the strength of the ash layer instead becomes crucial for the growth of
the roots (Paper III).
Generally a penetrometer resistance of 2-3 MPa is considered to prevent
plant root growth (Swedish Environmental Protection Agency, 1999). However, additional factors like soil particle size and structure are important;
values ranging between 0.8 MPa and 5.0 MPa have been found to restrict
root growth (Bengough and Mullins, 1990). In Paper IV, the aim was to find
out which penetration resistance is needed in a sealing layer of fly ash to
prevent root penetration. Although the ability to penetrate the fly ash layer
varied with the plant species, roots were in general prevented if the mechanical resistance exceeded ~1.5 MPa (Paper IV). Nevertheless, small amounts
of roots were found, especially of the grass Phalaris arundinacea, in the
surface layer of ash layers with resistances of ~5 MPa. These roots seemed
to have grown into pores or cracks in the fly ash, which is a common manner
of root growth in compacted soils (Dexter, 1986; Nicoullaud et al., 1994).
When a root meets compacted soil, and cannot change direction, it will decrease its growth rate by reducing both production and elongation of its cells
(Bennie, 1991). Simultaneously, the width of cortex cells will increase (Bengough et al., 1997), and thereby also the root diameter (Bennie, 1991; Materechera et al., 1991). Roots of P. arundinacea found in fly ash often had
increased diameters, and demonstrated coiled growth into pores (Paper IV).
By increasing its turgor pressure, the root can grow and penetrate compacted
soils (Materechera et al., 1991). Such forces exerted by the grass roots may
have been involved in loosening the fly ash, since in many cases the upper
33
surface of well-cured fly ashes was partly pulverized where roots were present (Paper IV).
From the field observations, it was clear that the formation of an impermeable sealing layer is not limited to fly ash with a high curing ability (Paper IV). A soil profile study of the dry cover at the test plots at Boliden mine
tailings impoundments showed that the fly ash had not cured well three years
after application. However, in the ~80 cm deep sealing layer, various layers
had formed, with a cementitious hard pan layer 5 to 15 cm below the ash
surface. No roots were found growing through this layer, but roots were
found growing extensively in the porous ash layer above. In a parallel test
plot where sludge was mixed with fly ash in the sealing layer, no hard pans
were observed, and plant roots were found down to 25 cm in the sealing
layer (unpublished data). In a greenhouse study (Paper III), addition of
sludge to the fly ash also increased penetration of plant roots, probably because it increased the amount and availability (by lowering the pH) of nutrients (Paper III), and decreased the bulk density of the material. In other studies, the combination of fly ash with inorganic or organic nutrients has
formed substrates supporting plant growth (Pillman and Jusaitis, 1997;
Cheung et al., 2000). Thus, even though addition of sewage sludge to fly ash
in the sealing layer may have other beneficial effects, it should be avoided in
order to prevent root penetration of the layer.
4.2.3 Plant loosening effects on cured fly ash
Fly ash exposed to outdoor conditions will start weathering: minerals, such
as ettringite and calcite, which are important for the strength of the fly ash
may dissolve (Steenari et al., 1999). Carbon dioxide in the air has a documented effect on fly ash, causing decreased pH and decalcification (Ecke,
2003). Since plants can both affect pH and release carbon dioxide from their
roots, they theoretically could be able to affect the stability of a fly ash layer.
When Phalaris arundinacea had been growing for 8 months on top of hardened fly ash, roots had been able to grow down into the fly ash and loosened
the surface of the ash (Paper IV). The same was observed by Greger et al.
(2006) in a greenhouse experiment lasting for 1 year, with various plant species growing in sewage sludge on top of cured fly ash. In that study a porous
surface of the ash was found only below roots of P. arundinacea but not of
Betula pubescens, Epilobium angustifolium, Picea abies, Pinus sylvestris,
Salix myrsinifolia, or Tussilago farfara. Analysis by XRPD (X-ray powder
diffraction) of the loosened fly ash in Paper IV indicated that the roots had
caused weathering of some minerals. While the content of the highly resistant mineral quartz was higher in rhizosphere ash than in bulk ash, the opposite was true for calcite (in biofuel fly ash) and gypsum (in MSW fly ash)
(Paper IV). This effect is possibly connected with the ability of roots to re34
duce pH (Paper III); the roots may have speeded up a process that otherwise
would have occurred abiotically. However, the effect can also be a result of
exuded substances that changed the mineralogy of the fly ash.
Roots exude substances as a response toward various environmental conditions (Neumann and Römheld, 2000). To decrease the frictional resistance
between root and soil, root cap cells are continuously produced and sloughed
off (Bengough et al., 1997), and the exudation of carbohydrates from the
root has been shown to increase with increasing mechanical impedance
(Barber and Gunn, 1974; Groleau-Renaud et al., 1998). Analysis of saccharide content in root exudates from P. arundinacea growing toward ash layers
of various strengths, however, did not reveal any such response (Paper IV).
Instead the exudation of saccharides seemed to increase if the roots were
exposed to a metal-rich fly ash as compared with a more alkaline fly ash.
The composition of root exudates is affected by many factors, including
nutrient availability (Marschner, 1995). Weathering of minerals can be significantly increased by the activity of growing roots, which has been connected with changes in elemental concentrations and pH, as well as exudation of chelating substances (Hinsinger et al., 2001). Dissolution of minerals,
in particular secondarily formed minerals, is often associated with the presence of organic acids (Hinsinger et al., 2001; Ryan et al., 2001; Reichard et
al., 2005). Organic acids found in manure have been found to exert a dissolving effect on cement (Bertron et al., 2005). The penetration resistance of
hardened fly ash was, however, not influenced by addition of a mixture of
organic acids, but was affected by addition of a mixture of monosaccharides
(fructose and glucose) (Paper IV). When ash has been mixed with cellulose,
iso-saccharinic acid has been formed (a degradation product of cellulose),
and increased the mobility of metals (Wikman et al., 2003). Furthermore,
saccharides are known to retard the curing of concrete (Garci Juenger and
Jennings, 2002), and can prevent formation of ettringite, possibly by chelating Ca (Cody et al., 2004). Since the fly ashes we tested were cured for only
one week before the solutions were added (Paper IV), they probably were
still reactive, and the saccharides may have prevented the growth of crystals.
However, the effect of the saccharides may also have been indirect, since the
solutions were not sterile; microorganisms may have used the saccharides
and exuded other compounds affecting the fly ash (Fig. 14). Microorganisms
can cause dissolution of minerals in a number of ways (Sand, 1997), and are
attracted to root surfaces, especially following increased root exudations, for
instance due to growth in compacted soils (Ikeda et al., 1997).
Further studies under axenic conditions could highlight the effects of
different substances on cured fly ash; however, since field conditions are
never sterile, it is important to understand the effects of complex systems
with both roots and associated microorganisms.
35
Root
CO2
Pressure
Fly ash
H+
Men+
Mineral
dissolution
Microorganisms
Root
exudates
e.g.
LMWOA
Carbo-hydrates
?
Figure 14. Schematic illustration
of possible impacts of plant roots
on cured fly ash. LMWOA= low
molecular weight organic acids.
4.3 Long term function and choice of plants
The time frame of the studies presented in this thesis extended over the
course of days to several months, complemented by observations in the field
of vegetated dry covers with ages of up to eight years (Paper IV). The results
should, therefore, be used for assessing short-term performance of vegetated
covers containing fly ash and sewage sludge.
The leaching of nutrients and metals will probably cease with time; however, the duration of this process is difficult to predict. In the greenhouse
study (Paper I), the release of nitrogen decreased within 2.5 months. In the
field study (unpublished data), there was a trend toward decreasing levels of
ammonium and phosphate, but not of nitrate, in sludge pore water during the
summer seasons (Fig. 11).
Construction of the dry cover is of great importance. Whether sludge or
fly ash is placed in direct contact with the tailings seems, however, not to be
as important as the compactness of the sealing layers. Layers that are too
shallow and porous will increase the risk of desiccation, crack formation,
root penetration and oxygen diffusion into the mine tailings. The depths of
sealing layers used in the field tests (0.5 to 0.8 m) seemed to be sufficient to
prevent root penetration. The fly ash used should have robust curing ability;
however, precipitation processes under outdoor conditions may cause cementitious layers even in fly ashes with weak curing ability (Paper IV).
Therefore, such ashes should not be disregarded until more results concerning their long-term function have been obtained. The protective cover of
36
sludge should be at least 0.5 m, since application of a thinner layer may desiccate before plants have established properly, and the layer may lose its
protective function. The sludge will be degraded with time, but will eventually be replaced by organic matter from the vegetation (Bendfeldt et al.,
2001). Since trees will disperse to the area, sooner or later, the protective
cover needs to be thick enough to support roots of large trees (Table 2).
Scars in the cover caused by trees felled by the wind may occur, but if the
ash sealing layer is compact enough to prevent root growth, it should also
remain intact even after a local removal of the protective cover; with time a
new organic layer may form on that spot. To keep the ash sealing layer resistant over a long period of time and prevent root growth, no or little sludge
should be added, as observed previously in a study of waste product degradation (Fang and Wong, 2000).
The plant species used in the experiments displayed varying ability to establish in sewage sludge, to prevent nutrient and metal leakage and to grow
into ash sealing layers (Table 6).
Table 6. Experiences from experiments with various plant species. Where there is no
reference, data is presented in the summary of the thesis.
Plant species
Grasses
Agrostis capillaris
Phalaris arundinacea
Poa pratensis
Experimental observations
Paper
Low germination rate in sludge
Vigorous growth in sludge, high ability to prevent
nutrient leakage and to grow into ash layers
Low germination rate in sludge, moderate growth
into ash layers
I, II, III,
IV
III
Herbs
Cannabis sativa
Vigorous growth in sludge and shallow root sysIII
tems
Epilobium angus- Vigorous growth in sludge, high ability to prevent I, IV
tifolium
nutrient leakage, little root growth into ash layers
Phaseolus vulgaris Sensitive toward ashes rich in metals and salts
III
Pisum sativum
Sensitive toward alkaline ashes
Tussilago farfara Vigorous growth in sludge
Trees and shrubs
Betula pendula
Low germination rate in sludge. Low root growth III
into ash layers
Picea abies
Medium growth in sludge
Pinus sylvestris
Low ability to prevent elemental leakage from
I, III, IV
sludge, and moderate growth into ash layers
Salix myrsinifolia Vigorous growth in sludge, low growth into ash
IV
layers
Salix viminalis
Sensitive toward fresh sludge. Moderate ability to I, III
prevent elemental leakage from sludge. High
growth into ash layers
37
Favorable plant species for a rapid establishment in sewage sludge are those
that can be dispersed by seeds. Plants spread by natural succession may have
a preventive effect on leakage (e.g., E. angustifolium), but will become established more slowly than a manually introduced plant. The energy grass P.
arundinacea exhibited vigorous growth in sludge in a temperate climate over
several seasons and was efficient in reducing leakage of both nutrients and
metals and is, therefore, a good option. However, this plant demonstrated a
strong ability to grow into sealing layers of fly ash (Paper IV). The protective layer must, therefore, be deep enough to support the root system of this
plant, and trees that will establish later on. Alternately, fast growing grass
species with shallower root systems should be chosen. Salix species may be
less suitable, since some of the species exhibit high metal uptake to the
shoots and restricted ability to prevent the release of nutrients and metals, at
least in the short term (Paper I). An introduction of various plant species
would, on the other hand, be beneficial, since they then can complement
each other with varying efficiencies in the uptake of different nutrients, as
discussed in Paper I and by Marschner (1998).
38
5 Conclusions
The general conclusion from the work presented in this thesis is that certain
plants may be efficient in phytostabilization of mine tailings covered with fly
ash and sewage sludge. Fast-growing plant species showed good potential in
reducing the leakage of N, P and metals. However, some plants were physiologically stressed, could not decrease the leakage from the sewage sludge, or
had an impact on the mineral structure of the sealing layer. The correct
choice of plant species is, therefore, of great significance. Moreover, the
sealing layer needs to be compact enough to prevent the penetration of plant
roots, and the protective cover deep enough to provide enough space as well
as nutrients and water for the root systems to develop. More specific conclusions, related to the aim of the thesis, are as follows:
1.
Establishment of plants in sewage sludge and their effect on leakage
of nutrients and metals.
a. A range of plant species can successfully establish in sewage
sludge even in a temperate climate. Ruderate plant species with
wind-dispersed seeds, such as Epilobium angustifolium and Salix
spp., grew vigorously within one year in areas left for natural succession. Among the introduced plants, seed-dispersed grasses
were more efficiently established than perennial plants that were
transferred as cuttings. However, in fresh sludge most plants did
not germinate, probably due to high levels of ammonium and
salts. Aeration of the sludge or addition of bark improved the
early establishment.
b. Plants decreased leaching of nutrients and metals mainly by decreasing the amount of leachate formed, but also by decreasing
the concentrations of elements. A high water uptake rate resulted
in reduced leakage of most elements. However, if roots were able
to grow through the sealing layer of fly ash down to the mine tailings, high water uptake caused desiccation of the mine tailings
and enhanced metal release. Plants with an ability to take up N
both as nitrate and ammonium (Phalaris arundinacea and Epilobium angustifolium) were more efficient in preventing N leakage
than those taking up primarily ammonium (Salix viminalis and
39
Pinus sylvestris). Fast-growing plants could neutralize pH in
leachate from sludge. However, the initial drop in pH and high N
leaching could not be counteracted by any of the plants used, and
plants did not affect pore water chemistry in field experiments.
Instead, seasonal variations had a larger impact.
c. Metals primarily accumulated in the roots of plants growing in
sewage sludge, while the translocation to shoots was moderate.
Hence, the risk to grazing animals seemed negligible. However,
shoots of S. myrsinifolia contained levels of Zn exceeding the
threshold levels for animal feed.
2.
Root penetration of sealing layers containing fly ash.
a. Plants respond negatively toward fresh fly ash, due to a combination of factors. Alkalinity strongly inhibits growth in many plants,
while less alkaline ashes induced stress responses in plants due to
high metal and salt content. In ash extractions with pH values up
to 12, plants could decrease the pH, provided that the alkalinity
was not too strong.
b. When the reactivity of the ashes decreases, root growth is instead
inhibited by high mechanical impedance of the material. Plant
roots were in general prevented from growing into an ash sealing
layer if the penetration resistance exceeded ~1.5 MPa. However,
small amounts of roots, in particular of P. arundinacea, grew into
the surface of ash layers with higher penetration resistances. Addition of sludge to the fly ash increased the risk of root penetration. Preferably, fly ash with strong curing ability should be chosen when constructing the sealing layer. Nevertheless, ashes with
weak curing ability can prevent root growth, if cementitious layers (hard pans) are formed in the sealing layer.
c. Root growth had an impact on the mineralogy of cured fly ash
and increased the weathering rate of secondary minerals, such as
calcite and gypsum. Exudation of organic compounds, such as
monosaccharides, may have an impact on the penetration resistance of a fly ash sealing layer.
40
6 Future work
To optimize the use of plants in the stabilization of mine tailings covered
with sewage sludge and fly ash, the physiological responses to these materials should be further examined. The exudation of substances from roots by
plants confronted with a sealing layer of fly ash should be investigated, with
regard to the composition of different ashes, as well as the plant species used
in the remediation and their effects on the ash layer. Furthermore, the effects
of plant roots on already-weathered ash layers, or degraded sewage sludge,
should be investigated, as well as the ability of different plant species to
utilize nutrients present in the sludge.
The investigations performed within the research for this thesis have in
general been focused on more short-term effects, which are those discovered
during the first months after application of a cover material and the establishment of plants. Most work has, furthermore, been conducted under
greenhouse conditions with plants growing in monocultures in small containers. To verify the data obtained, long-term studies should be performed,
with larger lysimeters or in the field. Further, leakage water should be collected from complete systems, containing all cover materials, as well as
vegetation, at points where the leachate is of concern.
Additional knowledge about the root penetration of old sealing layers is
necessary if we are to fully understand the effects of different types of vegetation on dry cover performance. The purpose of the dry cover is to encapsulate the mine tailings for as long as possible. During plant succession, trees
will establish and soil-forming processes will occur, rendering long-term
studies necessary.
41
Sammanfattning på svenska
Växter har en förmåga att anpassa och etablera sig i de flesta typer av jordar,
och kan också påverka förhållandena i jorden. På många platser används
växter för att stabilisera jord som annars skulle erodera, eller för att
immobilisera, ta upp eller bryta ner föroreningar, metoder som gemensamt
kallas för fytoremediering. Med en ökande medvetenhet om vad
föroreningar kan orsaka ökar också kraven på att industriella restprodukter
tas om hand på ett hållbart sätt, samtidigt som produktionen av avfall ökar.
Ett exempel på avfall som kräver stora deponeringsytor är sulfidrik
anrikningssand från utvinning av metaller. Denna finkorniga sand måste
täckas över för att inte börja vittra och släppa ifrån sig ett surt metallhaltigt
läckagevatten som är giftigt för många organismer. Eftersom deponierna är
stora krävs enorma mängder täckningsmaterial, och man utvärderar därför
möjligheterna att använda andra restprodukter för detta syfte.
Två produkter som också produceras i stora mängder och kräver
omhändertagande är flygaska från värmeverk och rötslam från
avloppsreningsverk. Flygaska har fördelen att det kan härda ungefär som
cement, och har låg nedbrytningshastighet, och därför skulle kunna forma ett
hållbart tätskikt som förhindrar syre och även växter och djur att ta sig ned
och påverka gruvavfallet. De flesta flygaskor har dessutom ett högt pH,
vilket skulle kunna motverka bildning av surt läckagevatten. Rötslam har
länge delvis använts inom jordbruket på grund av sitt höga näringsinnehåll,
men en rädsla för överföring av ohälsosamma kemikalier till livsmedel har
gjort att denna användning har minskat. På gruvavfall skulle slammet kunna
blandas med flygaska till ett tätskikt, eller läggas ut som skyddsskikt ovanpå
tätskiktet, som substrat för växtetablering. För att skapa ett täckskikt med
lång hållbarhet bör dock förhållandevis tjocka skikt anläggas, med följden att
stora mängder aska och slam sprids ut. Frågan är vad som händer när växter
etablerar sig – hur kan de påverka läckage av näring och metaller från
slammet, och kommer rötter att kunna växa ned i ett tätskikt av härdad aska?
Dessa är huvudfrågorna för denna avhandling, och svar har sökts genom ett
antal försök både i fält, växthus och laboratorium.
Resultaten visar att växter kan etablera sig väl i ett täckskikt av rötslam,
även om groning i färskt slam kan vara problematiskt. Växterna kan
förhindra läckage av föroreningar, men också öka läckaget om tätskiktet är
för tunt. Effekten varierade även mellan olika växtarter. Det initiala läckaget
av kväve visade sig kunna vara högt från rötslammet, men kan dämpas
42
framförallt om snabbväxande växer används med en god förmåga att ta upp
nitrat, ett näringsämne som är mycket lättlösligt. Framförallt kunde växter
förhindra läckage genom att ta upp vatten, och därmed minska mängden
bildat läckagevatten. Efter en initial insådd av gräs bör ett behandlat område
kunna lämnas för naturlig succession som tillåter att en blandning av
växtarter, i och med att olika arter visade potential att minska läckage av
olika ämnen, och kan därför komplettera varandra. Introduktion av
snabbväxande energigrödor är ett möjligt alternativ. Energigräset rörflen
förhindrade effektivt läckage av näring och metaller, men hade också en
förmåga att tolerera och växa ned i härdad flygaska, och kräver därmed ett
tjockt täckskikt. Korgvide, som också används som energigröda (”Salix”),
hade sämre förmåga att etablera sig i slammet och förhindra tidigt läckage av
kväve.
Nyhärdad flygaska är reaktiv, och kan förhindra rotväxt genom högt pH,
och innehåll av metaller och salter. När askan har stabiliserats och pH sjunkit
är dock hårdheten i askan avgörande för om rötter växer ned eller inte. Ett
penetrationsmotstånd över ~1.5 MPa räckte i allmänhet för att stoppa rotväxt
ned i askan, men små mängder rötter återfanns i ytan av asklager med högre
hårdhet. Rötterna, eller till dem associerade mikroorganismer, kunde
påskynda vittringen av mineral i askan, och anses därmed kunna påverka
hållbarheten i askan. Denna effekt är förmodligen dock liten jämfört med
vad abiotiska faktorer orsakar ute i fält. Observationer i fält visade att två
asktätskikt effektivt förhindrade rotväxt på olika sätt. Den ena askan hade
härdat mycket väl, och stoppade totalt all rotväxt även av välväxta träd. Den
andra askan hade härdat dåligt, men istället hade ett cementerat skikt bildats
i övre delen av tätskiktet som rötter inte kunde tränga igenom.
Sammanfattningsvis visar försöken att valda växter kan användas vid
stabilisering av gruvavfall med flygaska och rötslam, förutsatt att materialen
läggs ut så att rötter förhindras att växa ned i gruvavfallet.
43
Acknowledgements
This project had not been possible to carry out without support from a great
number of people. All of you that have been around me, giving me inspiration, discussed the project and supported me in many ways – thank you!
Especially, I want to thank Maria Greger, my supervisor, for years of encouragement, guidance, and interesting inputs. Lena Kautsky – you have
always been there as a trustful, calm and intelligent co-supervisor. I am also
most grateful for comments on the thesis from Birgitta Bergman, and good
advice from many other people at the department, e.g., Katharina Pawlowski, and Patrick Dinnetz.
All the former and present people in the Metal Group – you have made
this work inspiring, by discussing, helping, and playing heavy metal…
Thank you Agneta, Ann Helén, Beata, Claes, Eva, Johanna, Kathrin, Lisa,
Lovisa, Sylvia, Tariq, Tommy, Yaodong, and Åsa. Tommy – thanks for all
hours of laboratory expertise and interesting thoughts you have shared with
me. Johanna – what would I have done without your company, you have
always been able to cheer me up, and to come with clever input whenever it
was needed. Eva – thanks for all help when I started, and Yaodong – you
were such a nice person to collaborate with! I am also thankful to all other
collegues at the physiology unit, and at the department, for many hours of
nice company. Not to forget the people at the innebandy court – you made
one hour a week really sweaty and fun to me!
In the greenhouse I have gotten not only a lot of help, but also many interesting chats about plants with Peter and Ingela. Hasse – you really can fix
everything… and Ahmed – thanks for creating such a nice atmosphere in the
house.
At other universities, I want to acknowledge all those who have helped
me with physical analyses of ashes: in particular Andreas Fischer (KTH),
Dan Boström (Umeå University), Johan Arvidsson (SLU), and Per Delin
(KTH).
I am most grateful to Professor Jaco Vangronsveld at Limbergs University Centre in Belgium for welcoming me to his lab, and to the nice people in
the group for taking care of me during my stay.
In connection with all the travel to the field experiments in the north, I
have received much help from Karl-Erik Isaksson, Jenny Gotthardsson,
Björn Bergmark and Maria Lundmark at Boliden Mineral AB. Thank you
44
also, Anders Lindström at Högskolan Dalarna, for giving us access to the
field plots at Garpenberg.
The financial support should not be forgotten, and has kindly been given
by Värmeforsk (within the research program ”Environmentally friendly uses
of none coal ashes”), The Swedish Water & Waste Water Association (VAforsk), Boliden Mineral, Swedish Environmental Protection Agency, and
Stockholm Vatten. I am also thankful towards the thermal power stations in
Hässelby, Högdalen, Iggesund, Munksund, Hedensbyn, Fors, and Umeå, for
providing me with fly ashes. Wiking Pettersson, Frikonsult, the initiator of
the project, Claes Ribbing and the reference group of the “ash project”
(Värmeforsk) – thanks for all your enthusiasm!
I have also been lucky with having a number of fabulous persons around
me outside work: Mum and dad, thanks for being fantastic, always so motivating! David and Love for not only being brothers, but also friends. IngMarie and Carl-Otto, I am so happy for all your assistance, and inspired by
your positive attitude. Then there are my close friends – you have maybe
wondered what on earth I am doing, but you have always been there, lightening up my life – thank you! I have also been most encouraged by you people
in the volleyball team, and in the dinner group.
Last but not least comes the greatest thank from my heart to Peter, you are
like an engine – always coming with bright ideas, questioning any too simple
“truth”, and just being warm and supportive. And Matteus, you wonderful
little guy, thanks for your love!
45
References
Abad, M., Noguera, P. and Burés, S. 2001. National inventory of organic wastes for
use as growing media for ornamental potted plant production: case study in
Spain. Bioresource Technol. 77:197-200.
Adriano, D.C., Page, A.L., Elseewi, A.A., Chang, A.C. and Straughan, I. 1980.
Utilization and disposal of fly ash and other coal residues in terrestrial ecosystems: A review. J. Environ. Qual. 9:333-344.
Antoniadis, V., Robinson, J.S. and Alloway, B.J. 2008. Effects of short-term pH
fluctuations on cadmium, nickel, lead, and zinc availability to ryegrass in a sewage sludge-amended field. Chemosphere 71:759-764.
Arvidsson, J. 1999. Nutrient uptake and growth of barley as affected by soil compaction. Plant Soil 208:9-19.
Bakker, M.R. and Nys, C. 1999. Effect of Liming on Fine Root Cation Exchange
Sites of Oak. J. Plant Nutr. 22:1567-1575.
Barber, D.A. and Gunn, K.B. 1974. The effect of mechanical forces on the exudation of organic substances by the roots of cereal plants grown under sterile conditions. New Phytol. 73:39-45.
Basta, N.T., Ryan, J.A. and Chaney, R.L. 2005. Trace Element Chemistry in Residual-Treated Soil: Key Concepts and Metal Bioavailability. J. Environ. Qual.
34:49-63.
Bell, L.C. 2001. Establishment of native ecosystems after mining – Australian experience across diverse biogeographic zones. Ecol. Eng. 17:179-186.
Bendfeldt, E.S., Burger, J.A. and Daniels, W.L. 2001. Quality of Amended Mine
Soils After Sixteen Years. Soil Sci. Soc. Am. J. 65:1736-1744.
Bengough, A.G., Croser, C. and Pritchard, J. 1997. A biophysical analysis of root
growth under mechanical stress. Plant Soil 189:155-164.
Bengough, A.G. and Mullins, C.E. 1990. Mechanical impedance to root growth: a
review of experimental techniques and root growth responses. J. Soil Sci.
41:341-358.
Bennie, A.T.P. 1991. Growth and Mechanical Impedance. In: Waisel, Y. and Eshel,
A. (eds) Plant Roots – The Hidden Half. Marcel Deccer Inc, New York, USA,
pp 393-416.
Bertron, A., Duchesne, J. and Escadeillas, G. 2005. Attack of cement pastes exposed
to organic acids in manure. Cement Concrete Comp. 27:898-909.
Brady, N.C. and Weil R.R. 2002. The Nature and Properties of Soils (13th ed).
Prentice Hall, Upper Saddle River, NJ.
Bramryd, T. 2002. Impact of sewage sludge application on the long-term nutrient
balance in acid soils of Scots pine (Pinus sylvestris. L) forests. Water Air Soil
Poll. 140:381-399.
Brown, S.L., Henry, C.L., Chaney, R., Compton, H. and deVolder, P.S. 2003. Using
municipal biosolids in combination with other residuals to restore metalcontaminated mining areas. Plant Soil 249:203-215.
46
CEC (Council of the European Communities). 1991a. Council Directive of 21 May
1991 concerning urban waste-water treatment (91/271/EEC). Off. J. European
Communities No. L 135.
CEC (Council of the European Communities). 1991b. Council Directive of 31 Dec.
1991 concerning the protection of waters against pollution caused by nitrates
from agricultural sources (91/676/EEC). Off. J. European Communities No. L
371/1.
Cheung, K.C., Wong, J.P.K., Zhang, Z.Q., Wong, J.W.C. and Wong, M.H. 2000.
Revegetation of lagoon ash using the legume species Acacia auriculiformis and
Leucaena leucocephala. Environ. Poll. 109:75-82.
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 Agriculural Sciences. Report 47.
Clesceri, L.S. and Greenberg, A.E. (eds). 1995. Standard Methods for the Examination of Water and Wastewater. Eaton, A. D.
Cody, A.M., Lee, H., Cody, R.D. and Spry, P.G. 2004. The effects of chemical environment on the nucleation, growth, and stability of ettringite
[Ca3Al(OH)6]2(SO4)3·26H2O. Cement Concrete Res. 34:869-881.
Cravotta III, C.A. 1998. Effect of Sewage Sludge on Formation of Acidic Ground
Water at a Reclaimed Coal Mine. Ground Water 35:9-19.
Demeyer, A., Voundi Nkana, J.C. and Verloo, M.G. 2001. Characteristics of wood
ash and influence on soil properties and nutrient uptake: an overview. Bioresource Technol. 77:287-295.
Dexter, A.R. 1986. Model experiments on the behaviour of roots at the interface
between a tilled seed-bed and a compacted sub-soil. III. Entry of pea and wheat
roots into cylindrical biopores. Plant Soil 95:149-161.
Ecke, H. 2003. Sequestration of metals in carbonated municipal solid waste incineration (MSWI) fly ash. Waste Manage. 23:631-640.
Eckstein, R.L. and Karlsson, P.S. 1997. Above-ground growth and nutrient use by
plants in a subarctic environment: effects of habitat, life-form and species. Oikos
79:311–324.
Elander, P., Lindvall, M. and Håkansson, K. 1998. MiMi – Prevention and control
of pollution from mining waste products. State-of-the-art-report. MiMi 1998:2
Eriksson, J. 2001. Halter av 61 spårelement i avloppsslam, stallgödsel,
handelsgödsel, nederbörd samt i jord och gröda. Report 5148. (In Swedish),
Swedish Environmental Protection Agency, Stockholm, Sweden.
Evanylo, G.K., Abaye, A.O., Dundas, C., Zipper, C.E., Lemus, R., Sukkariyah, B.
and Rockett, J. 2005. Herbaceous Vegetation Productivity, Persistence, and
Metals Uptake on a Biosolids-Amended Mine Soil. J. Environ. Qual. 34:18111819.
Fang, M. and Wong, J.W.C. 2000. Changes in thermophilic bacteria population and
diversity during composting of coal fly ash and sewage sludge. Water Air Soil
Poll. 124:333-343.
Fjällborg, B. and Dave, G. 2003. Toxicity of copper in sewage sludge. Environ. Int.
28:761-769.
Fjällborg, B., Ahlberg, G., Nilsson, E. and Dave, G. 2005. Identification of metal
toxicity in sewage sludge leachate. Environ. Int. 31:25-31.
Fröberg, G. and Höglund, L.O. 2004. MiMi Light – en populärvetenskaplig
sammanfattning av MiMi-programmets forskning kring efterbehandling av
gruvavfall. MiMi Report 2004:8. (In Swedish). MiMi Print, Luleå, Sweden.
47
Fuentes, A., Lloréns, M., Sáez, J., Aguilar, M.I., Pérez-Marín, A.G., Ortuño, J.F. and
Mesetuer, V.F. 2006. Ecotoxicity, phytotoxicity and extractability of heavy
metals from different stabilised sewage sludges. Environ. Poll. 143:355-360.
Fytili D. and Zabaniotou A. 2008. Utilization of sewage sludge in EU application of
old and new methods—A review. Renew. Sust. Energ. Rev. 12: 116-140.
Garcia-Gomez, A., Bernal, M.P. and Roig, A. 2002. Growth of ornamental plants in
two composts prepared from agroindustrial wastes. Bioresource Technol. 83:8187.
Garci Juenger, M.C. and Jennings, H.M. 2002. New insights into the effects of sugar
on the hydration and microstructure of cement pastes Cement Concrete Res.
32:393-399.
Gibert, O., de Pablo, J., Cortina, J.L. and Ayora, C. 2004. Chemical characterisation
of natural organic substrates for biological mitigation of acid mine drainage.
Water Res. 38:4186-4196.
Gigon, A. and Rorison, I. H. 1972. The response of some ecologically distinct plant
species to nitrate- and to ammonium-nitrogen. J. Ecol. 60: 93-102.
Giordano, P.M., Behel, A.D. Jr., Lawrence, J.E.Jr., Solieau, J.M. and Bradford, B.N.
1983. Mobility in soil and plant availability of metals derived from incinerated
municipal refuse. Environ. Sci. Technol. 17:193-198.
Gomez, E., Durillon, C., Rofes, G. and Picot, B. 1999. Phosphate adsorption and
release from sediments of brackish lagoons: pH, O2 and loading influence. Water Res. 33:2437-2447.
Greger, M., Neuschütz C. and Isaksson, K-E. 2006. Flygaska och rötslam som
tätskikt vid efterbehandling av sandmagasin med vegetationsetablering. Report
959 (In Swedish). Värmeforsk, Stockholm, Sweden.
Greger, M. and Landberg, T. 2008. Role of rhizosphere mechanisms in Cd uptake by
various wheat cultivars. Plant Soil 312:195-205.
Gregory, P.J. 2006. Roots, rhizosphere and soil: the route to a better understanding
of soil science? Eur. J. Soil Sci. 57:2-12.
Groleau-Renaud, V., Plantureuz, S. and Guckert, A. 1998. Influence of plant morphology on root exudation of maize subjected to mechanical impedance in hydroponic conditions. Plant Soil 201:231-239.
Guo, F-Q., Wang, R., Chen, M. and Crawford, N. M. 2001. The Arabidopsis dualaffinity nitrate transporter gene AtNRT1.1 (CHL1) is activated and functions in
nascent organ development during vegetative and reproductive growth. Plant
Cell 13:1761-1778.
Gupta, E.K., Rai, U.N., Tripathi, R.D. and Inouhe, M. 2002. Impacts of fly-ash on
soil and plant responses. J. Plant Res. 115:401-409.
Han, Z.H., Shen, T., Korcak, R.F. and Baligar, V.C. 1998. Iron absorption by ironefficient and –inefficient species of apples. J. Plant Nutr. 21:181-190.
Haynes, R.J. 1980. Ion exchange properties of roots and ionic interactions within the
root apoplasm: their role in ion accumulation by plants. Bot. Rev. 46:75-99.
Hearing, K.C., Daniels, L.W. and Feagley, S.E. 2000. Reclaiming mined lands with
biosolids, manures, and papermill sludges. In: Reclamation of Drastically Disturbed Lands, Agronomy Monograph no. 41. American Society of Agronomy,
Crop Science Society of America, Madison, US.
Heintze, S.G. 1961. Studies on cation-exchange capacities of roots. Plant Soil
13:365-383.
Hinsinger, P., Fernandes Barros, O.N., Benedetti, M.F., Noack, Y. and Callot, G.
2001. Plant-induced weathering of a basaltic rock: Experimental evidence. Geochim. Cosmochim. Ac. 65:137-152.
48
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.
Holmström, H. 2000. Geochemical processes in sulphidic mine tailings. Doctoral
thesis. Department of Environmental Engineering, Division of Applied Geology, Luleå University of Technology, Sweden.
Holmström, H., Salmon, U.J., Carlsson, E., Petrov, P. and Öhlander, B. 2001. Geochemical investigations of sulfide-bearing tailings at Kristineberg, northern
Sweden, a few years after remediation. Sci. Tot. Environ. 273:111-133.
Holtzclaw, K.M., Keech, D.A., Page, A.L., Sposito, G., Ganje, T.J. and Ball, N.B.
1978. Trace Metal Distributions among the Humic Acid, The Fulvic Acid, and
Precipitable Fractions Extracted with HaOH from Sewage Sludges. J. Environ.
Qual. 7:124-127.
Ikeda, K., Toyota, K. and Kimura, M. 1997. Effects of soil compaction on the microbial populations of melon and maize rhizoplane. Plant Soil 189:91-96.
Kumpiene, J., Lagerkvist, A. and Maurice, C. 2007. Stabilization of Pb- and Cucontaminated soil using coal fly ash and peat. Environ. Poll. 145:365-373.
Larcher, W. 2003. Physiological Plant Ecology. (4th ed). Springer. Berlin.
Larsson, I., Maripuu, R., Isaksson, K-E. and Lindvall, M. 2005. The new tailings
storage facility in Boliden– a pioneering project for mill tailings management in
the 21:st century. Securing the Future – International Conference on Mining and
the Environment, June 27 – July 1 2005, Skellefteå, Sweden. Conference Proceedings, 557-565.
Ledin, M. and Pedersen, K. 1996. The environmental impact of mine wastes – Roles
of micro-organisms and their significance in treatment of mine wastes. EarthSci. Rev. 41:67-108.
Mácsik, J., Rogbeck, Y., Svedberg, B., Uhlander, O. and Mossakowska, A. 2003.
Fly ash and sewage sludge as liner material – Preparation for a pilot test with
fly-ash stabilized sewage sludge as landfill liner. Report 837. (In Swedish).
Värmeforsk, Stockholm, Sweden.
Madoni, P. and Romeo, M.G. 2006. Acute toxicity of heavy metals toward freshwater ciliated protists. Environ. Poll. 141:1-7.
Marschner, H. 1995. Mineral Nutrition of Higher Plants. (2nd ed.). Academic Press,
London, UK.
Marschner, H. 1998. Role of root growth, arbuscular mycorrhiza, and root exudates
for the efficiency in nutrient acquisition. Field Crop. Res. 56:203-207.
Materechera, S.A., Dexter, A.R. and Alston, A.M. 1991. Penetration of very strong
soils by seedling roots of different plant species. Plant Soil 135:31-41.
Merrington, G., Oliver, I., Smernik, R.J. and McLaughlin, M.J. 2003. The influence
of sewage sludge properties on sludge-borne metal availability. Adv. Environ.
Res. 8:21-36.
Miller, R.W., Al-Khazraji, M.L., Sisson, D.R. and Gardiner, D.T. 1995. Alfalfa
growth and absorption of cadmium and zinc from soils amended with sewage
sludge. Agr. Ecosyst. Environ. 53:179-184.
Min, X., Siddiqi, M.Y., Guy, R.D., Glass, A.D.M. and Kronzucker, H.J. 2000. A
comparative kinetic analysis of nitrate and ammonium influx in two earlysuccessional tree species of temperate and boreal forest ecosystems. Plant Cell
Environ. 23:321-328.
Misra, R.K., Dexter A.R. and Alston, A.M. 1986. Penetration of soil aggregates of
finite size. I. Blunt penetrometer probes. Plant Soil 94:43-58.
Mollamahmutoglu, M. and Yilmaz, Y. 2001. Potential use of fly ash and bentonite
49
mixture as liner or cover at waste disposal areas. Environ. Geol. 40:1316-1324.
Murphy, J. and Riley, J.P. 1962. A modified single-solution method for the determination of phosphate in natural waters. Anal. Chim. Ac. 27:31-36.
Nasr, H.M. and Selles, F. 1995. Seedling emergence as influenced by aggregate size,
bulk density, and penetration resistance of the seed bed. Soil Till. Res. 34:61-76.
Naylor, L.M. and Schmidt, E.J. 1986. Agricultural use of wood ash as a fertilizer
and liming material. Tappi J. 69:114-119.
Neumann, G. and Römheld, V. 2000. The release of root exudates as affected by the
plant physiological status. In: Pinton, R., Varanini, Z. and Nannipieri, Z. (eds.)
The Rhizosphere: Biochemistry and organic substances at the soil-plant interface. Marcel Dekker, New York, US.
Nicoullaud, B., King, D. and Tardieu, F. 1994. Vertical distribution of maize roots
in relation to permanent soil characteristics. Plant Soil 159:245-254.
Nordin, A., Högberg, P. and Näsholm, T. 2001. Soil nitrogen form and plant nitrogen uptake along a boreal forest productivity gradient. Oecologia 129: 125-132.
Notter, M. 1993. Metallerna och miljön (The metals and the environment). Report
4135. (In Swedish). Swedish Environmental Protection Agency, Stockholm,
Sweden.
NRC (National Research Council). 2005. Mineral Tolerance of Animals (2nd Rev.
Ed.). National Academies Press, Washington DC, USA.
Pais, I. and Jones, J.B.Jr. 2000. The Handbook of Trace Elements. St. Lucie Press,
Boca Raton, Florida, US, p 50.
Peppas, A., Komnitsas, K. and Halikia, I. 2000. Use of organic covers for acid mine
drainage control. Min. Eng. 13:563-574.
Pérez-López, R., Nieto, J.M. and Almodóvar, G.R. 2007. Utilization of fly ash to
improve the quality of the acid mine drainage generated by oxidation of a sulphide-rich mining waste: Column experiments. Chemosphere 67:1637-1646.
Pichtel, J. and Anderson, M. 1997. Trace metal bioavailability in municipal solid
waste and sewage sludge composts. Bioresource Technol. 60:223-229.
Pillman, A. and Jusaitis, M. 1997. Revegetation of waste fly ash lagoons II. Seedling
transplants and plant nutrition. Waste Manag. Res. 15:359-370.
Potters, G., Pasternak, T.P., Guisez, Y., Palme, K.J. and Jansen M.A.K. 2007.
Stress-induced morphogenic responses: growing out of trouble? Trends Plant
Sci. 12:98-105.
Prashanth, J.P., Sivapullaiah, P.V. and Sridharan, A. 2001. Pozzolanic fly ash as a
hydraulic barrier in land fills. Eng. Geol. 60:245-252.
Rate, A.W., Lee K.M. and French P.A. 2004. Application of biosolids in mineral
sands mine rehabilitation: use of stockpiled topsoil decreases trace element uptake by plants. Bioresource Technol. 91:223-231.
Reichard, P.U., Kraemer, S.M., Frazier, S.W. and Kretzschmar, R. 2005. Goethite
dissolution in the presence of phytosiderophores: rates, mechanisms, and the
synergistic effect of oxalate. Plant Soil 276:115-132.
Ribbing, C. 2007. Environmentally friendly use of non-coal ashes in Sweden. Waste
Manage. 27:1428-1435.
Ribet, I., Ptacek, C.J., Blowes, D.W. and Jambor, J.L. 1995. The potential for metal
release by reductive dissolution of weathered mine tailings. J. Contam. Hydrol.
17:239-273.
Rimstidt, J.D. and Vaughan, D.J. 2003. Pyrite oxidation: A state-of-the-art assessment of the reaction mechanism. Geochim. Cosmochim. Ac. 67: 873-880.
50
Ryan, P.R., Delhaize, E. and Jones, D.L. 2001. Function and Mechanism of Organic
Anion Exudation from Plant Roots. Ann. Rev. Plant Physiol. Plant Mol. Biol.
52:527-60.
Sajwan, K.S., Paramasivam, S., Alva, A.K., Adriano, D.C. and Hooda, P.S. 2003.
Assessing the feasibility of land application of fly ash, sewage sludge and their
mixtures. Adv. Environ. Res. 8:77-91.
Salt, D.E., Blaylock, M., Kumar, N.P.B.A., Dusenkov, 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. Biotechnol. 13:468-474.
Sand, W. 1997. Microbial Mechanisms of Deterioration of Inorganic Substrates – A
General Mechanistic Overview. Int. Biodeter. Biodegr. 40:183-190.
Santibáñez, C., Ginocchio, R. and Varnero, M.T. 2007. Evaluation of nitrate leaching from mine tailings amended with biosolids under Mediterranean type climate conditions. Soil Biol. Biochem. 39:1333-1340.
Santibáñez, C., Verdugo, C. and Ginocchio, R. 2008. Phytostabilization of copper
mine tailings with biosolids: Implications for metal uptake and productivity of
Lolium perenne. Sci. Tot. Environ. 395:1-10.
Sastre, J., Sahuquillo, A., Vidal, M. and Rauret, G. 2002. Determination of Cd, Cu,
Pb and Zn in environmental samples: microwave-assisted total digestion versus
aqua regia and nitric acid extraction Anal. Chim. Ac. 462:59-72.
Scandiaconsult Sverige AB. 2001. Slam i mark- och anläggningsbyggande.
Avvattnat vattenverks- och avloppsslam. Report 1, 2001. (In Swedish), Stockholm Vatten, Sweden.
Scheetz, B.E. and Earle, R. 1998. Utilization of fly ash. Curr. Opin. Solid St. M.
3:510-520.
Schenk, H.J. and Jackson, R.B. 2002. The global biogeography of roots. Ecol.
Monogr. 72:311-328.
Sivapullaiah, P.V., Prashanth, J.P., Sridharan, A. and Narayana, B.V. 1998. Technical note: Reactive silica and strength of fly ashes. Geotechn. Geol. Eng. 16:239250.
Smith, R.A.H. and Bradshaw, A.D. 1979. The use of metal tolerant plant populations for the reclamation of metalliferous wastes. J. Appl.Ecol. 16, 595-612.
Smith, M.T.E. and Tibbett, M. 2004. Nitrogen dynamics under Lolium perenne after
a single application of three different biosolids types from the same treatment
stream. Bioresource Technol. 91:233-241.
Sopper, W.E. 1993. Municipal Sludge Use in Land Reclamation. Lewis Publishers,
CRC Press, Boca Raton, US.
Statistics Sweden. 2008. Discharges to water and sewage sludge production in
2006. Published in collaboration with the Swedish Environmental Protection
Agency, at: http://www.scb.se/templates/Publikation____232141.asp.
Steenari, B-M. and Lindqvist, O. 1997. Stabilisation of biofuel ashes for recycling to
forest soil. Biomass Bioenerg. 13:39-50.
Steenari, B-M., Karlsson, L.G. and Lindqvist, O. 1999. Evaluation of the leaching
characteristics of wood ash and the influence of ash agglomeration. Biomass
Bioenerg. 16:119-136.
Stehouwer, R., Day, R.L. and Macneal, K.E. 2006. Nutrient and trace element leaching following mine reclamation with biosolids. J. Environ. Qual. 35:1118-1126.
Stirzaker, R.J., Passioura, J.B. and Wilms, Y. 1996. Soil structure and plant growth:
Impact of bulk density and biopores. Plant Soil 185:151-162.
51
Stjernman Forsberg, L. 2008. Reclamation of Copper Mine Tailings Using Sewage
Sludge. Doctoral Thesis, Department of Soil and Environment, Swedish University of Agriculture, Uppsala, Sweden.
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.
Stouraiti, C., Xenidis, A. and Paspaliaris, I. 2002. Reduction of Pb, Zn and Cd availability from tailings and contaminated soils by the application of lignite fly ash.
Water Air Soil Poll. 137:247-265.
Stuczynski, T., Siebielec, G., Daniels, W.L., McCarty, G. and Chaney, R.L. 2007.
Biological Aspects of Metal Waste Reclamation with Biosolids. J. Environ.
Qual. 36:1154-1162.
Su, D.C. and Wong, J.W.C. 2003. Chemical speciation and phytoavailability of Zn,
Cu, Ni and Cd in soil amended with fly ash-stabilized sewage sludge. Environ.
Int. 29:895-900.
Swedish Environmental Protection Agency (SEPA). 1999. Environmental Quality
Criteria for Agricultural Landscapes (In Swedish). Report 4916, SEPA, Stockholm, Sweden.
Södermark, B. 1986. Gruvavfall. Naturvårdsverket informerar. Brochure (In Swedish), 11 p, Swedish Environmental Protection Agency. Stockholm, Sweden.
Tahvanainen, L. and Rytkönen, V-M. 1999. Biomass production of Salix viminalis
in southern Finland and the effect of soil properties and climate conditions on its
production and survival. Biomass Bioenerg. 16:103-117.
Tham, G. and Andreas, L. 2008. Results from a full scale application of ashes and
other residuals in the final cover construction of the Tveta landfill. Report 1064
(In Swedish). Värmeforsk, Stockholm, Sweden.
Theodoratos, P., Moirou, A., Xenidis, A. and Paspaliaris, I. 2000. The use of municipal sewage sludge for the stabilization of soil contaminated by mining activities. J. Hazard. Mater. B77:177-191.
Topaç, F.O., Başkaya, H.S. and Alkan, U. 2008. The effects of fly ash incorporation
on some available nutrient contents of wastewater sludges. Bioresource Technol. 99:1057-1065.
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.
Tripathi, R.D., Vajpayee, P., Singh, N., Rai, U.N., Kumar, A., Ali, M.B., Kumar, B.
and Yunus, M. 2004. Efficacy of various amendments for amelioration of flyash toxicity: growth performance and metal composition of Cassia siamea
Lamk. Chemosphere 54:1581-1588.
USDA, NRCS. 2009. The PLANTS Database (http://plants.usda.gov, 3 January
2009). National Plant Data Center, Baton Rouge, LA 70874-4490 USA.
Van Assche, F. and Clijsters, H. 1990. A biological test system for the evaluation of
the phytotoxicity of metal-contaminated soils. Environ. Poll. 66:157-172.
Vaz, C.M.P., Bassoi, L.H. and Hopmans, J.W. 2001. Contribution of water content
and bulk density to field soil penetration resistance as measured by a combined
cone penetrometer-TDR probe. Soil Till. Res. 60:35-42.
Vedin, H. 2005. Sveriges landskapsklimat: Dalarnas klimat. (In Swedish). SMHI,
Swedish Meteorological and Hydrological Institute. Väder och Vatten 10: 10.
Vedin, H. 2007. Sveriges landskapsklimat: Västerbottens klimat. (In Swedish).
SMHI, Swedish Meteorological and Hydrological Institute. Väder och Vatten 8:
10.
52
Venendaal, R., Jørgensen, U. and Foster, C.A. 1997. European energy crops: A
synthesis. Biomass Bioenerg. 13:147-185.
Villar, L.D. and Garcia Jr.,O. 2002. Solubilization profiles of metal ions from
bioleaching of sewage sludge as a function of pH. Biotechnol. Lett. 24:611-614.
WEF. 2009. A Guide to Understanding Biosolids Issues. (http://www.wef.org 13
Jan. 2009), Water Environment Federation, Alexandria, USA.
Wei, X., Roger C. Viadero Jr., R.C. and Bhojappa, S. 2008. Phosphorus removal by
acid mine drainage sludge from secondary effluents of municipal wastewater
treatment plants. Water Res. 42:3275-3284.
Weng, L., Van Riemsdijk, W.H. and Hiemstra, T. 2007. Adsorption of humic acids
onto goethite: Effects of molar mass, pH and ionic strength . J. Colloid Interf.
Sci. 314:1-118.
Westermark, M. and Gromulski, J. 1996. Termisk kadmiumrening av
trädbränsleaskor. Ramprogram askåterföring. Report 1996:30 (In Swedish).
NUTEK, Vattenfall, Sydkraft, Stockholm, Sweden.
Wikman, K., Berg, M., Svensson, M. and Ecke, H. 2003. Degradation of cellulose in
the presence of ash. Report 806 (In Swedish). Värmeforsk, Stockholm, Sweden.
Wiles, C.C. 1996. Municipal solid waste combustion ash: State-of-the-knowledge. J.
Hazard. Mater. 47:325-344.
Wong, J.W.C., Jiang, R.F. and Su, D.C. 1998. The accumulation of boron in Agropyron elongatum grown in coal fly ash and sewage sludge mixture. Water Air
Soil Poll. 106:137-147.
Wong, M.H. 2003. Ecological restoration of mine degraded soils, with emphasis on
metal contaminated soils. Chemosphere 50:775-780.
Xie, Z. and Xi, Y. 2001 Hardening mechanisms of an alkaline-activated class F fly
ash. Cement Concrete Res. 31:1245-1249.
Ye, Z.H., Yang, Z.Y., Chan, G.Y. 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. 26:449-55.
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