Arsenic transformations in the soil rhizosphere plant system fundamental and potential application to ph
Trang 1Arsenic transformations in the soil/rhizosphere/plant system: fundamentals and potential application to phytoremediation
Walter J Fitz, Walter W Wenzel * Institute of Soil Science, University of Agricultural Sciences Vienna-BOKU, Gregor Mendel Strasse 33, A-1180 Vienna, Austria
Received 3 September 2001; received in revised form 24 May 2002; accepted 27 May 2002
Abstract
This paper reviews major processes that potentially affect the fate of arsenic in the rhizosphere of plants Rhizosphere interactions are deemed to play a key role in controlling bioavailability to crop plants and for a better understanding and improvement of phytoremediation technologies Substantial progress has been made towards an understanding of
As transformation processes in soils However, virtually no information is available that directly addresses the fate of
integrating the state-of-the art knowledge available in the contributing disciplines Using this model and recent studies
on hyperaccumulation of As, we discuss research needs and the potential application of rhizosphere processes to the development of phytoremediation technologies for As-polluted soils
Keywords: Arsenic; Bioavailability; Hyperaccumulation; Mycorrhiza; Phytoremediation; Rhizosphere
1 Introduction
Arsenic is an ubiquitous trace metalloid and is
found in virtually all environmental media
How-ever, concentrations of As in non-contaminated
soils are typically well below 10 mg kg1 Its
presence at elevated concentrations in soils is due
to both anthropogenic and natural inputs
Anthro-pogenic sources include mining and smelting
processes besides application of As-based
insecti-cides, herbiinsecti-cides, fungiinsecti-cides, algiinsecti-cides, sheep dips,
wood preservatives, dyestuffs, feed additives and compounds for the eradication of tapeworm in sheep and cattle (Adriano, 2001) Geochemical sources of As-contaminated soils include As-rich parent material as As easily substitutes for Si, Al
or Fe in silicate minerals (Bhumbla and Keefer,
1994) Arsenic is also commonly associated with sulfides, e.g in sulfidic ore deposits Other natural sources of As include volcanic activities, wind-born soil particles, sea salt sprays and microbial volatilisation of As (Nriagu, 1990; Frankenberger and Arshad, 2002)
It has been estimated that there are potentially 1.4 million contaminated sites within the European Community impacted to various extent by organic and/or trace metal/metalloid pollutants (European
* Corresponding author Tel.: /43-1-47654-3119; fax:
/43-1-4789-110
E-mail address: wwenzel@edv1.boku.ac.at (W.W Wenzel).
0168-1656/02/$ - see front matter # 2002 Elsevier Science B.V All rights reserved.
PII: S 0 1 6 8 - 1 6 5 6 ( 0 2 ) 0 0 2 1 8 - 3
Trang 2Topic Centre Soil, 1998) Forty-one percent of the
superfund sites in the USA for which US EPA has
signed records of decision are contaminated with
As (US EPA, 1997), more than 10 000
As-con-taminated sites have been reported for Australia
(Smith et al., 2002) Though considerable progress
has been made in reducing atmospheric inputs of
As in Western Europe (Schulte and Gehrmann,
1996), pollution by As and other trace metals at a
large scale can still occur as shown by the Don˜ana
ecological disaster in southern Spain (Pain et al.,
1998)
Drinking of As-contaminated groundwater is
perhaps the most common exposure pathway of
humans to As toxicity The biggest known As
calamity occurred in the Bengal Delta
(Bangla-desh/West Bengal) where millions of people
de-pend on As-rich drinking water (Chakraborti et
al., 2001) Natural contamination of groundwater
by As has been also recorded for many other parts
in the world.Berg et al (2001)reported a recently
discovered case of groundwater contamination in
Hanoi (Vietnam) with contamination levels
vary-ing from 1 to 3050 mg l1
Technologies currently available for the
reme-diation of metal/metalloid contaminated soils are
expensive, time consuming, can create risks to
workers and produce secondary waste (Wenzel et
al., 1999a; Lombi et al., 2000a,b) Recently
phy-toremediation, the use of green plants to clean up
contaminated soil, has attracted much attention
(Baker et al., 1991; McGrath et al., 1993)
Basi-cally, the strategy of phytoremediation can be
divided into five fundamental processes that apply
to As, including phytoextraction, stabilisation,
immobilisation, volatilisation and rhizofiltration
(Salt et al., 1998; Wenzel et al., 1999a,b) A major
step towards the development of
phytoremedia-tion of As-impacted soils is the recent discovery of
the As-hyperaccumulating ferns Pteris vittata and
Pityrogramma calomelanos Both plants produce
large biomass and are therefore promising
candi-dates for phytoextraction purposes (Ma et al.,
2001; Francesconi et al., 2002; Tu and Ma, 2002;
Visoottiviseth et al., 2002)
This paper is focused on the basic processes
involved in As transformation in the soil /
rhizosphere plant system Rhizosphere
interac-tions are deemed to play a key role in controlling bioavailability to crop plants (Hinsinger, 2001) and for a better understanding and improvement
of phytoremediation technologies (Wenzel et al., 1999b; Lombi et al., 2001) However, virtually no literature is available which refers particularly to the biogeochemistry of As in the rhizosphere The bulk of available literature is related to more general aspects of soil /plant /As relationships Several comprehensive reviews are available on
As in the soil system (e.g Bhumbla and Keefer 1994; O’Neill, 1995; Sadiq, 1997; Smith et al., 1998; Adriano, 2001) However, the fate of As in the rhizosphere has yet not been explored This review addresses major processes potentially in-volved in the fate and transformation of As in the soil /rhizosphere /plant system in order to present conceptual models and hypotheses while high-lighting future research needs to enhance the scientific basis for further development of phytor-emediation technologies
2 Arsenic transformations in the soil/plant/ microbe system
2.1 General Arsenic has been known to have a high affinity for oxide surfaces, which is affected by several biogeochemical factors such as soil texture, or-ganic matter, nature and constituents of minerals,
pH, redox potential and competing ions (Adriano,
2001) The activity of As in soil solution is most commonly controlled by surface complexation reactions on oxides/hydroxides of Al, Mn and especially Fe (Inskeep et al., 2002) Smaller textural fractions contain larger sorbed and total amounts of As (Lombi et al., 2000a,b), as sorbing oxides/hydroxides are typically concentrated in the clay size fraction ( B/2 mm) due their small size This explains also the lower toxicity of fine textured compared with coarse textured As-pol-luted soils (Jacobs et al., 1970) Analyses of drainage waters derived from mine tailings have shown that suspended material ( /0.45 mm) is the main carrier of arsenic and mainly responsible for
Trang 3metal fluxes into ground and surface waters
(Roussel et al., 2000)
About 25 different As compounds have been
identified in biological samples, mainly in marine
ecosystems (Francesconi and Edmonds, 1993)
However, usually only the organic species
mono-methylarsonic acid (MMAA) and dimethylarsinic
acid (DMAA) are found in detectable
concentra-tions in soils besides abundant inorganic AsV and
AsIIIspecies (Takamatsu et al., 1982,Tlustosˇ et al.,
2002) Paddy soils typically show larger
extracta-ble concentrations of MMAA and DMAA which
suggests that methylated arsenicals are produced
under anaerobic conditions (Takamatsu et al.,
1982) In very few cases trimethylarsine oxide
(TMAO) and arsenobetaine (AB) have been
de-tected as minor compounds in soil extracts (
Geis-zinger et al., 2002)
Toxicity and chemical behaviour of As
com-pounds are largely influenced by the form and
speciation of As AsIII is more mobile and more
toxic than AsV Gaseous arsines are most toxic
whereas arsenobetaine and arsenocholine (mainly
found in marine organisms) are nontoxic As a
rule, inorganic arsenicals are more toxic than
organic arsenicals and the trivalent oxidation state
is more toxic than the pentavalent oxidation state
(Fowler, 1977; Adriano, 2001)
Though most studies did not directly investigate
the fate of As in the rhizosphere we highlight in the
following the major processes taking place in the
rhizosphere to assess the potential interactions
with the fate of As in the soil /plant /microbe
system
2.2 Fate of arsenic as related to rhizosphere
acidification/alkalinisation
It is generally known that rhizosphere pH may
considerably differ from that in the bulk soil
Depending on plant and soil factors pH
differ-ences can be up to two units Factors affecting
rhizosphere pH are the source of nitrogen supply
(NO3 vs NH4 uptake), nutritional status of
plants (e.g Fe and P deficiency), excretion of
organic acids, CO2 production by roots and
rhizosphere microorganisms, and the buffering
capacity of the soil (Marschner, 1995)
Several studies have been carried out on pH-dependent As sorption in soils and on pure mineral phases Studies using soil and pure Fe hydroxides generally agree that AsV solubility increases upon pH increase within pH-ranges commonly found in soil (pH 3 /8), whereas AsIII tends to follow the opposite pattern (Manning and Goldberg, 1997; Smith et al., 1999; Tyler and Olsson, 2001; Raven et al., 1998; Jain and Loep-pert, 2000) Thermodynamic calculations suggest that H2AsO4 dominates below and HAsO42
above pH 6.97 (Sadiq, 1997) Furthermore, net surface charges of soil constituents become more negative as functional groups dissociate protons upon pH increase Conversion of H2AsO4 to HAsO42 along with increasing negative surface charges of soil constituents lead to AsV mobilisa-tion as electrostatic repulsion is enhanced particu-larly above pH 7 ( /pK2) Moreover, as well the oxide concentration of soil has considerable influ-ence on the pH-dependent solubility of As.Smith
et al (1999) found that for soils low in oxidic minerals pH had little effect on the amount of adsorbed AsVwhereas highly oxidic soils showed a pronounced decrease of AsVadsorption upon pH increase In contrast to AsV, solubility of AsIII decreases with decreasing pH in soil The pK1of arsenous acid (H3AsO3) is 9.22, which implies that below pH 9.22 the AsIII species is mainly uncharged (Sadiq, 1997) This may contribute to the generally larger solubility of AsIII in soil systems
Most soils exhibit oxic conditions, hence an increase of rhizosphere pH could favour mobilisa-tion of labile and exchangeable AsV-fractions in the root vicinity and consequently enhance plant uptake Nitrogen nutrition, as it is most respon-sible for the cation/anion uptake ratio, greatly affects rhizosphere pH (Marschner and Ro¨mheld,
1983) Hence, fertilisation of plants grown on As-contaminated soil with NO3 as the N source, would potentially increase rhizosphere pH, and thus possibly enhance As accumulation in plant tissues On the other hand there are distinct differences in rhizosphere acidification among plant species For instance legumes and actinorhi-zal plants meet their N supply by symbiontic N
Trang 4fixation N2 enters the root uncharged, thus the
cation/anion uptake ratio of N2 fixing plants is
large and results in a net H release by the plant
(Marschner, 1995) Rhizosphere acidification by
N2-fixing symbionts would favour AsV
immobili-sation in soil under oxic conditions
The As hyperaccumulator P vittata was
re-ported to prefer calcareous soils of neutral to
slightly alkaline pH (Jones, 1987; Ma et al., 2001)
This implies that changes of rhizosphere pH would
be no prerequisite for As-hyperaccumulation due
to the high pH-buffer power of calcareous soils
However, P vittata and P calomelanos have been
as well found on acidic soils and mine tailings in
Thailand An increase of the rhizosphere pH could
potentially increase AsV solubility and possibly
plant uptake on such substrates On the other
hand very low pH values may dissolve As sorbents
such as Fe oxides/hydroxides (seeSections 2.6 and
2.4)
2.3 Root exudation
It has been reported that P-deficient plants show
an enhanced exudation of carboxylic acids, such as
citric and malic acid (Hoffland, 1992; Neumann
and Ro¨mheld, 1999; Kirk et al., 1999) This
response is thought to change soil pH, to displace
P from sorption sites, to chelate metal cations that
could immobilise P or to form soluble
metal-chelate complexes with P, resulting in enhanced
availability of P (Kirk et al., 1999) Cluster roots of
P-deficient plant species such as Lupinus albus and
members of the Protaceae exude particularly
strong organic acids and phenolics (Dinkelaker et
al., 1995) Arsenic and P belong to the same
chemical group and have comparable dissociation
constants for their acids and solubility products
for their salts, resulting in similar geochemical
behaviour of As and P in soil (Adriano, 2001)
Hence, it is reasonable to assume that carboxylate
exudation could play a role in the mobilisation of
As in the rhizosphere and enhance As uptake by
plants
Basically two strategies have been identified for
acquisition of Fe by higher plants Strategy I exists
in monocotyledenous species, with the exception
of graminaceae (grasses), and dicotyledenous
spe-cies, and involves three processes: (1) enhanced net excretion of protons, (2) a plasma membrane-bound inducible reductase, and (3) enhanced release of reducing and chelating agents Strategy
II, confined to grasses, is characterised by release
of phytosiderophores and a high-affinity transport system for Fe uptake (Marschner and Ro¨mheld,
1994) Fe-oxides/hydroxides typically dominate As sorption in soil (see Section 2.7) Laboratory studies of arsenate and arsenite adsorption on Fe-oxide surfaces indicate that both species are bound as mono and bidentate surface complexes (Waychunas et al., 1993; Sun and Donner, 1996) The excretion of protons and/or the release of reducing and chelating compounds by strategy I plants also could result in co-dissolution of As from Fe-oxides/hydroxides, rendering As more soluble and available to plants
Admittedly, virtually nothing is known about
Fe nutritional aspects and related rhizosphere processes of fern plants They are sensu strico neither strategy I nor strategy II plants as ferns belong to the Pteridophyta However, ferns such the As-hyperaccumulator P vittata and P calo-melanos certainly acquire Fe P vittata is known
to grow commonly on calcareous soil (Jones,
1987) It has been reported that root exudates (oxalic and citric acid) of acidifuge plants effec-tively mobilise P and Fe from lime stone (Stro¨m et al., 1994) Porter and Peterson (1975) found a highly significant correlation (P B/0.001) between
As and Fe in several As-tolerant plants from different mine sites in UK No correletions were found between As and other elements (Pb, Cu, Zn), not even for P
In conclusion we suggest that P, Fe and As uptake by As hyperaccumulator species may be related to each other Reductive dissolution of
FeIII minerals inevitably dissolves Fe-bound As, root exudates enhancing P mobilisation are likely to desorb As as well Besides rhizosphere processes As-hyperaccumulator most likely posses
a particular As-uptake mechanism whereas sup-pression of the high-affinity phosphate uptake system is involved in adaptive tolerance of plants
to As (Meharg and Macnair, 1992a see Section 2.7)
Trang 52.4 Redox potential
The redox potential significantly influences
speciation and solubility of As in soils (e.g.Deuel
and Swoboda, 1971; Masschelyn et al., 1991;
Marin et al., 1993; McGeehan and Naylor, 1994;
Onken and Hossner, 1995, 1996) Generally,
inorganic As is mainly present as AsV in aerobic
conditions (high redox potential) and as AsIII in
flooded (low redox potential) soils Arsenic is less
toxic and less mobile in the /V than in the /III
oxidation state It has been repeatedly observed
that increased As solubility under reduced
condi-tions is associated with dissolution of Fe and Mn
oxides/hydroxides Significant correlations have
been found between dissolved Fe and As
(Masschelyn et al., 1991; Marin et al., 1993;
McGeehan and Naylor, 1994), confirming that
Fe oxides/hydroxides represent the major sorbing
agents for As in soils (see Section 2.6) Flooding
had no influence on soluble Ca and Al (
Massche-lyn et al., 1991).Masschelyn et al (1991)
investi-gated redox /pH relations of AsV and AsIII
stability using an apparatus which allowed pH
and redox control of a stirred soil suspension
Under oxidised conditions, soluble As
concentra-tions were three times larger at pH 8 than at pH 5,
because of the decreased positive surface charge at
high pH Under reducing conditions AsIIIbecame
the major dissolved species with total soluble As
being smaller at pH 8 Dissolved Fe concentrations
did not significantly increase upon reduction at pH
8 (Masschelyn et al., 1991) In contrast,Marin et
al (1993), using the same experimental set up,
reported increased As solubility upon pH decrease
(7.5 /5.5) for both reduced and oxidised conditions
without providing any explanation As
concentra-tions in rice (Oryza sativa L.) increased upon
decreasing redox potential (Marin et al., 1993)
The oxidation of the rhizosphere is a well known
phenomenon for paddy rice as these plants are
able to transport O2through aerenchyma to roots,
which results in a leakage of O2 into the
rhizo-sphere (Flessa and Fischer, 1992) Rice roots
grown in reduced suspensions were coated with
Fe plaque containing As (Marin et al., 1993)
Doyle and Otte (1997) found formation of Fe
plaque also around roots of salt marsh plants
which led to an effective fixation and consequently detoxification of As and Zn in the rhizosphere 2.5 As /P interactions
Similar to carboxylic acids released by plant roots, other organic and inorganic anions may compete with As for sorption sites The phosphate ion plays a prominent role in anion /As interac-tions due to its physicochemical similarity to As (Adriano, 2001) Moreover, arsenate is thought to
be taken up via the phosphate uptake system and may consequently interact with plant P nutrition (Asher and Reay, 1979; Meharg and Macnair,
1990) Though numerous studies on As /P inter-actions have been published, results have not been explored systematically and yet have not been applied to the rhizosphere Table 1 gives a compilation of studies on As /P interactions with respect to mobilisation/extractability, plant uptake and phytotoxicity of As Generally it is reasonable
to distinguish between hydroponics (solution cul-ture), pot/column/batch and field experiments Hydroponic experiments inevitably overestimate the importance of uptake kinetics of the plant in consideration (Meharg et al., 1994) and typical processes of soil /plant relationships such as water flow, nutrient/pollutant mass flow to the root surface, diffusion, adsorption/desorption and ion exchange are not considered as soil is absent in such an experimental set up Consequently, P additions in solution culture studies decrease As uptake and mitigate As-caused phytotoxicity symptoms (Hurd-Karrer, 1939; Asher and Reay, 1979; Tsutsumi, 1982; Meharg and Macnair, 1990;
De Koe and Jaques, 1993)
Summarising results of pot, laboratory and field experiments leads to a different conclusion Phos-phorus additions at high rates enhance As leaching
in laboratory column studies (Woolson et al., 1973; Peryea and Kammereck, 1995), increase extractable fractions of As in batch experiments (Carrow et al., 1975; Peryea, 1991) and reduced sorption of AsVand AsIIIonto soils (Smith et al.,
2002)
Plant uptake of As has been shown to increase upon P application in pot experiments (Creger and Peryea, 1994; Jiang and Singh, 1994; Woolson,
Trang 61972, Woolson et al., 1973) and at field scale
(Small and McCants, 1962) In contrast to solution
culture studies, presence of P causes As /P
compe-tition for sorption sites resulting in increased As
bioavailability, and hence higher As
concentra-tions in plants Quaghebeur and Rengel (2001)
studied As /P interactions in the rhizosphere and
found that the presence of P significantly increased
As concentrations in shoots and roots of both
tolerant and non-tolerant clones of Holcus lanatus
Jacobs and Keeney (1970)compared As
accumu-lation in corn from artificially contaminated soils
(20 and 80 mg kg1) Arsenic concentrations were
larger in plants grown on sandy soil compared
with a silty loam Increasing the level of P in soil
had little effect on plant uptake of As on the silt
loam but showed a marked increase on the sandy
soil when As was present at 80 mg kg1
Both reduced and increased phytotoxicity
symp-toms had been found after P additions to soil
grown plants Differences in texture and mineral
content affect also As /P relationships
Hurd-Karrer (1939)found in pot experiments improved
growth of As-injured wheat on clay loam and
sandy loam soils upon addition of P In contrast,
Woolson et al (1973)reported reduced growth of corn after P fertilisation on a sandy loam and enhanced growth on silty clay loam Jacobs and Keeney (1970)found enhanced As-toxicity to corn
on a sandy soil but little effects on a silty loam
Benson (1953) found good response of barley growth to added P only on seven of 17 soils with toxic concentrations of As Schweizer (1967)
studied toxicity of disodium methanearsonate (DSMA) residue applied on two silty loam soils and found that P additions increased phytotoxicity symptoms Benson (1953)tested P additions also
at field scale and found no yield response In this study superphosphate fertiliser was applied in dry form in 10 cm deep trenches 10 cm away from the seeds However, phosphorus is known to be very immobile in soils (Marschner, 1995), which likely resulted in spatially confined As /P /root interac-tions
In case of As-hyperaccumulating plants it is very unlikely that P fertilisation may cause phytotoxi-city problems as it has been reported that P vittata accumulates up to 22 630 mg As kg1 in dry matter on soil spiked with 1500 mg As kg1 (Ma et al., 2001) Therefore, P additions may in
Table 1
Selection of literature on the response of mobilisation, plant uptake, phytotoxicity of As to P additions in hydroponics, pot/column/ batch and field experiments
Hydroponic Pot/column/batch Field
Phytotoxicity
As III Slightly increased 9
a Disodium methanearsonate.
1, Peryea and Kammereck (1995) ; 2, Peryea (1991) ; 3, Woolson et al (1973) ; 4, Hurd-Karrer (1939) ; 5, Asher and Reay (1979) ; 6,
Carrow et al (1975) ; 7, Schweizer (1967) ; 8, Benson (1953) ; 9, Tsutsumi (1982) ; 10, Meharg and Macnair (1990) ; 11, De Koe and Jaques (1993) ; 12, Jiang and Singh (1994) ; 13, Jacobs and Keeney (1970) ; 14, Woolson (1972) ; 15, Small and McCants (1962) ; 16,
Creger and Peryea (1994) ; 17, Quaghebeur and Rengel (2001) ; 18, Smith et al (2002) ; 19, Sneller et al (1999)
Trang 7the first place enhance plant growth and secondly
mobilise exchangeable As resulting in increased
total As uptake
Most studies on As /P interactions were carried
out using spiked soil and delivered valuable
information However, concentrations of labile
As and P can be expected to be substantially
different in soils that gradually received As during
extended periods through various anthropogenic
processes such as mining and smelter activities
(Wenzel et al., 2002a) It has been demonstrated
that As rapidly becomes recalcitrant in soil with
time (Lombi et al., 1999; Onken and Adriano,
1997), resulting in reduced toxicity (Jiang and
Singh, 1994)
Large proportions of total soil P may be present
in organic forms such as phytates (Dalal, 1977;
Marschner, 1995) Similarly, up to 70% of
dis-solved P in soil solution were found to be present
as organic P (Helal and Sauerbeck, 1984) Phytic
acid is expected to compete with As for sorption
sites due to its anionic nature However,
interac-tions of As with organic P have not been studied
yet
2.6 Binding forms of As in soil
Despite apparent similarities between the
chem-istry of As and P, some important differences have
to be considered Unlike P, As is present also in
oxidation state III, and besides oxygen other
ligands may form stable species that are not found
with P (O’Neill, 1995)
The traditional Chang and Jackson (1957)
procedure, developed for sequential extraction of
P, has been adopted for fractionation of As in soils
(e.g.Woolson et al., 1971, 1973; Akins and Lewis,
1976; Onken and Hossner, 1996; Onken and
Adriano, 1997; Wasay et al., 2000) It has been
assumed that this extraction scheme addresses,
with respect to P, the so-called water-soluble plus
adsorbed (NH4Cl-extractable) As and the
Al-(NH4F-extractable), Fe- (NaOH-extractable),
and Ca-bound (H2SO4-extractable) As fractions
Based on this extraction procedure (Woolson et
al., 1971, 1973; Akins and Lewis, 1976; Wasay et
al., 2000) and comparisons between extractable
fractions of As and Fe/Al oxide/hydroxide content
of soils (Wauchope, 1975; Johnston and Barnard, 1979; Polemio et al., 1982; Manning and Gold-berg, 1997; Chen et al., 2002) it has been suggested that Fe-oxides/hydroxides represent the major sink for As sorption in soils, whereas the importance of Al- and Ca-bound fractions are variable In none
of the studies using a modifiedChang and Jackson (1957)procedure for As fractionation co-dissolved
Al, Fe and Ca were analysed in the extracts Studies presenting data on co-dissolved Al, Fe and Ca prove that the Ca-bound As plays a minor role in As sorption even in calcareous soils (Wenzel et al., 2001a; Shiowatana et al., 2001) Hence, only minor proportions of As were found
in extracts (1 M sodium acetate /acetic acid buffer)
of soils addressing Ca-bound metal fractions (Wenzel et al., 2001a) These findings are in agreement with results of energy dispersive X-ray microanalysis (EDXMA), providing evidence for strong association of As with hydrous Fe oxides (Lombi et al., 2000b) Oxides/hydroxides of Al,
Mn and Fe are also known to form coatings on other soil particles such as clays Fordham and Norrish (1983)reported that adsorption of AsVin
a lateritic Podzol was mainly controlled by Fe oxide deposits on kaolin flakes, showing a high degree of substitution of Al for Fe within Fe oxide particles
Little research has been done on As adsorption
by organic matter Thanabalasingam and Picker-ing (1986)found sorption of As onto humic acids
in batch experiments However, this was primarily related to the ash content of the humic acids used There is no evidence that soil organic matter (SOM) would contribute in significant quantities
to As sorption in soils, especially in the presence of effective sorbents such as hydrous Fe oxides (Livesey and Huang, 1981; Wenzel et al., 2001a) Risk assessment of As-polluted sites of both anthropogenic and geogenic origin even revealed enhanced As solubility in organic surface horizons
of forest soils (Brandstetter et al., 2000; Wenzel et al., 2002a) This can be explained by the anionic nature of many organic compounds in soil, result-ing in reduced As adsorption on Al and Fe oxides/ hydroxides (Fordham and Norrish, 1983; Xu et al., 1988)
Trang 8Turpeinen et al (1999) used a modified metal
sequential extraction procedure of Tessier et al
(1979)for As-fractionation They reported that up
to 14.4% of total As was organically bound,
however, without testing the procedure for its
applicability to As
Even though fractionation of soil-As by
sequen-tial extraction procedures delivers only
operation-ally defined As forms, its application in
rhizosphere investigations may be useful in order
to determine As pools of differential
bioavailabil-ity, e.g determination of As-pools potentially
accessible to hyperaccumulator plants Fig 1
shows range and median values (%) of As
dis-tribution among the five fractions of theWenzel et
al (2001a) sequential extraction procedure
Twenty polluted soils of both anthropogenic and
geogenic As origin were used These results reveal
that most As is associated with the Fe oxides/
hydroxides (fraction 3/4) The amount of the
non-specifically sorbed fraction (readily mobile) is
small but most important for risk assessment with
respect to potential ground water pollution
(Brandstetter et al., 2000; Wenzel et al., 2002a)
2.7 Plant uptake of As
There is no evidence that As is essential for
plants, though growth is stimulated when supplied
at low concentrations (Liebig et al., 1959; Lepp,
1981; Carbonell et al., 1998) From hydroponic
experiments on plant uptake of As it is known that
the chemical form of supplied As is more impor-tant than total As concentrations in solution In solution culture experiments Marin et al (1992)
found that the phytoavailability for two rice cultivars followed the order DMAA B/AsVB/ MMAA B/AsIII, whileCarbonell-Barachina et al (1998)obtained the order of DMAA B/MMAA $/
AsVB/AsIII for two typical wetland plant species
of the Lousiana salt marshes However, both reports agree that upon absorption, inorganic species and MMAA were mainly accumulated in roots In contrast DMAA was readily translocated
to the shoots resulting in shoot/root As concentra-tion ratios of /1.Tlustosˇ et al (2002)conducted
a pot experiment on As uptake in radish grown on soil amended with AsIII, AsV and DMAA AsIII was readily oxidised to AsV, resulting in no differences in As accumulation and yield between these two treatments Water extracts showed that DMAA was adsorbed to a much lesser extent than
AsV, causing a significant reduction of radish biomass production, although total As concentra-tions were similar to the other As treatments Plants capable of accumulating exceptionally large concentrations of metals have been termed hyperaccumulators (Brooks et al., 1977) Recently, the first As-hyperaccumulating plants, the ferns P vittata and P calomelanos have been discovered Both ferns produce large biomass, and are there-fore, promising candidates for phytoextraction purposes (Ma et al., 2001; Francesconi et al., 2002; Visoottiviseth et al., 2002) However, some confusion has entered the discussion on hyper-accumulator plants lately Formerly reported As-tolerant plants grown on heavily As-polluted soils and mine tailings have been repeatedly termed as
As hyperaccumulators (Francesconi et al., 2002; Francesconi and Kuehnelt, 2002; Visoottiviseth et al., 2002; Geiszinger et al., 2002).Table 2provides
an overview on reported shoot, root and substrate concentrations of As hyperaccumulators and As-tolerant plants The biological absorption coeffi-cients (BAC, defined as the total element concen-tration in shoots with respect to total element concentration in soil, both in mg kg1) and accumulation factors (AF, defined as the total element concentration in shoots with respect to total element concentration in roots, both in mg
Fig 1 Partitioning of As among the five fractions of a
sequential extraction procedure in 20 test soils Upper case
symbols refer to: (a) non-specifically sorbed, (b)
specifically-sorbed, (c) bound to amorphous and poorly-crystalline hydrous
oxides of Fe and Al, (d) bound to well-crystallised hydrous
oxides of Fe and Al Arsenic pollution was caused by both
natural and anthropogenic inputs Extracted from Wenzel et al.
(2001a)
Trang 9kg1) were calculated where possible
Compari-sons between hyperaccumulators and tolerant
plants evidently show the difference in the As
accumulation behaviour Whereas tolerant plants
tend to restrict soil /plant and root /shoot transfer,
hyperaccumulators actively take up and
translo-cate As into above-ground tissues Plants
exhibit-ing AF and particularly BAC values B/1 do not
represent candidates for phytoextraction P
vittata grown on As-spiked soil resulted in dry
matter concentrations of As as high as 22 630 mg
kg1 (Ma et al., 2001) Based on the
accumu-lator /excluder concept of Baker (1981) As
toler-ant pltoler-ants (Table 2) should be termed as excluders
at AF ratios /1, even they show elevated
concentrations in above ground tissues
Hyperaccumulation of As seems to be rather
constitutive than adaptive as populations from
non-contaminated environments hyperaccumulate
As as well (Table 2) Biomass production of P
vittata has been shown to increase upon As
applications, suggesting the status of a beneficial
element for this plant (Tu and Ma, 2002)
Root-induced rhizosphere processes can be anticipated
to facilitate As uptake by hyperaccumulator
plants
Arsenate is thought to be taken up via the phosphate uptake system (Asher and Reay, 1979; Meharg and Macnair, 1990) Solution culture studies revealed that tolerant populations of Agrostis capillaris (Porter and Peterson, 1975) and H lanatus (Meharg and Macnair, 1991a) take up less As than non-tolerant plants Meharg and Macnair (1992a)showed that arsenate uptake
in solution culture of a As-tolerant H lanatus population was caused by the suppression of the high affinity P uptake system (the high affinity uptake system for P is dominant at concentrations
B/0.1 mmol l1; Clarkson and Lu¨ttge, 1991), resulting in smaller As influx and accumulation
in tolerant populations Similar arsenate tolerance mechanisms were observed for As-tolerant popu-lations of Deschampsia cespitosa and to a lesser extent as well for A capillaris (Meharg and Macnair, 1991b) In contrast, no down regulation
of arsenate/phosphate transporters was found for As-tolerant Calluna vulgaris (Sharples et al., 2000a) Recent findings of Hartley-Whitaker et
al (2001) suggest that arsenate tolerance in H lanatus requires both adaptive suppression of the high-affinity phosphate uptake system and consti-tutive phytochelatin production Phytochelatins
Table 2
Arsenic accumulation in hyperaccumulator and As tolerant plants
Plant species As in plants (mg kg 1 ) As in soil (mg kg 1 ) BAC a AF b References
Frond/shoot Root Hyperaccumulators
Tolerant plants (non-accumulators)
Only data of live plant material was collected 1, Ma et al (2001) ; 2, Francesconi et al (2002) ; 3, Porter and Peterson, (1975) ; 4, De Koe, (1994) ; 5, Jonnalagadda and Nenzou (1996, 1997) ; 6, Bech et al (1997)
a Biological absorption coefficient (shoot/soil concentration ratio).
b Accumulation factor (shoot/root concentration ratio).
c Spiked soil.
Trang 10seem to be also involved in detoxification of As in
Silene vulgaris (Sneller et al., 1999) and Nicotiana
tabacum (Nakazawa et al., 2000) Most As in
fronds of P vittata and P calomelanos is present
as AsIII, whereas AsV dominates in roots (Ma et
al., 2001; Francesconi et al., 2002) As metabolism
and complexation in plants have been reviewed by
Meharg and Hartley-Whitaker (2002) and are
therefore not discussed here
From hydroponic experiments one may
con-clude that tolerant populations having a
sup-pressed phosphate/arsenate high affinity uptake
system are less efficient in absorbing P and hence
would produce less biomass.Meharg et al (1994)
compared P uptake of tolerant and non-tolerant
H lanatus populations in both solution culture
and in pot experiments with sterile potting
com-post At low P concentrations (0.5 and 5 mM) the
tolerant clones showed reduced plant P
concentra-tion and shoot biomass, but a higher percentage of
root biomass These differences were not found at
high P concentrations in solution (50 mM) In
contrast, tolerant plants grown in pots had smaller
growth rates but higher P concentrations in their
tissues As discussed by the authors, hydroponic
experiments tend to overestimate the importance
of uptake kinetics, as influx rather than P diffusion
(Nye, 1977), root morphology and soil parameters
(Silberbush and Barber, 1983) appear to be the
rate limiting steps
2.8 Mycorrhizal associations and other microbial
interactions in the rhizosphere
Mycorrhizas are the most widespread
mutualis-tic symbiomutualis-tic association between microorganisms
and higher plants and can be important for the
mineral nutrition of the host plant (Wilcox, 1991)
Apart from these well known beneficial effects on
plant nutrition mycorrhizal associations may fulfil
other functions for host plants growing on
con-taminated land Mycorrhizal fungi may alleviate
metal toxicity to the host plant by acting as a
barrier for metal uptake (Leyval et al., 1997)
Recently, Sharples et al (2000b) compared the
short-term uptake kinetics of the ericoid
mycor-rhizal fungus Hymenoscyphus ericae from an As/
Cu-contaminated mine site (As-resistant
popula-tion) and from an uncontaminated natural heath-land (non-As-resistant population) in solution culture Uptake kinetics of AsV, AsIII and phos-phate did not differ for resistant and non-resistant isolates However, the mine-site fungi showed an approximately 90% enhanced efflux of As in the form of AsIII Twenty-four-hours uptake of AsV
by hydroponically-grown mycorrhizal and non-mycorrhizal host C vulgaris did not differ for mine-site plants In contrast, inoculated heathland
C vulgaris accumulated 100% more As than non-inoculated individuals (Sharples et al., 2000a) The authors suggested that the mine site fungus acts as
an As filter to maintain low As concentrations in plant tissues, while improving P nutrition of the host plant Therefore, mycorrhizal fungi may be important for the revegetation/phytostabilisation
of As-polluted sites
Arsenic tolerance of H lanatus populations from non-contaminated sites was found to be polymorphic (Meharg and Macnair, 1992a,b)
Meharg et al (1994) investigated 50 tussocks from the same population of which 40% showed tolerance to As The As-tolerant phenotype had a 11% higher P-status and a 34% higher arbuscular mycorrhizal (AM)-infection rate of roots Wright
et al (2000) conducted a field experiment using clones of tolerant and non-tolerant H lanatus populations Though no difference in AM mycor-rhization could be observed, tolerant plants accu-mulated more P in shoots, had largerer shoot and root biomass and produced considerably more flower panicles Results of Meharg et al (1994), Wright et al (2000) show that conclusions drawn from studies on uptake kinetics in solution culture may have limited validity in more complex field conditions
Most ferns normally exhibit mycorrhizal asso-ciations (Jones, 1987) The role of mycorrhiza in
As hyperaccumulation is not known yet We found that P vittata individuals grown in pots were colonised by MA fungi Most well-studied hyper-accumulators belong to the Brassicaceae (Brooks, 1998; Baker et al., 2000), which generally do not form mycorrhizal associations (Marschner, 1995)
To our knowledge no studies on the role of any type of mycorrhizal symbiosis in hyperaccumula-tion has been carried out so far From present