Arsenic transformations in the soil rhizosphere plant system fundamental and potential application to ph

Arsenic transformations in the soilÁ/rhizosphereÁ/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
As in the rhizosphere. We are proposing a conceptual model of the fate of As in the soilÁ
/rhizosphereÁ/plant system by
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.
# 2002 Elsevier Science B.V. All rights reserved.
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 kg
(1
. 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, herbicides, fungicides, algicides, 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: (W.W. Wenzel).
Journal of Biotechnology 99 (2002) 259 Á
/278
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PII: S 0 1 6 8 - 1 6 5 6 ( 0 2 ) 0 0 2 1 8 - 3
Topic 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 mgl
(1
.
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 soilÁ
/
rhizosphereÁ/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 soilÁ
/plantÁ/As relationships.
Several comprehensivereviews 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
soilÁ
/rhizosphereÁ/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
Á
/plant
Á
/
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
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278260
metal fluxes into ground and surface waters
(Roussel et al., 2000).
About 25 different As compounds havebeen
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 As
V
and
As
III
species (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) havebeende-
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. As
III
is more mobile and more
toxic than As
V
. 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 soilÁ
/plantÁ/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
(NO
3
(

vs. NH
4
'
uptake), nutritional status of
plants (e.g. Fe and P deficiency), excretion of
organic acids, CO
2
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 As
V
solubility
increases upon pH increase within pH-ranges
commonly found in soil (pH 3Á
/8), whereas As
III
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 H
2
AsO
4
(
dominates below and HAsO

4
2(
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 H
2
AsO
4
(
to
HAsO
4
2(
along with increasing negative surface
charges of soil constituents lead to As
V
mobilisa-
tion as electrostatic repulsion is enhanced particu-
larly abovepH7(Â
/pK
2
). 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 As
V
whereas highly oxidic soils showed a

pronounced decrease of As
V
adsorption upon pH
increase. In contrast to As
V
, solubility of As
III
decreases with decreasing pH in soil. The pK
1
of
arsenous acid (H
3
AsO
3
0
) is 9.22, which implies
that below pH 9.22 the As
III
species is mainly
uncharged (Sadiq, 1997). This may contribute to
the generally larger solubility of As
III
in soil
systems.
Most soils exhibit oxic conditions, hence an
increase of rhizosphere pH could favour mobilisa-
tion of labile and exchangeable As
V
-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 NO
3
(
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
2
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278 261
fixation. N
2
enters the root uncharged, thus the
cation/anion uptake ratio of N
2
fixing plants is
large and results in a net H
'
release by the plant
(Marschner, 1995). Rhizosphere acidification by
N
2

-fixing symbionts would favour As
V
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 havebeen
as well found on acidic soils and mine tailings in
Thailand. An increase of the rhizosphere pH could
potentially increase As
V
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 (see Sections 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
¨
met
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
Fe
III
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).
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278262
2.4. 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 As
V
in aerobic
conditions (high redox potential) and as As
III
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 redoxÁ
/pH relations of As
V
and As
III
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 As
III
became
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
(7.5Á
/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 O
2
through aerenchyma to roots,
which results in a leakage of O
2
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. AsÁ
/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 Á
/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 Á
/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 AsÁ
/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 soilÁ
/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 As
V
and As
III
onto 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,
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278 263
1972, Woolson et al., 1973) and at field scale
(Small and McCants, 1962). In contrast to solution
culture studies, presence of P causes As Á
/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 Á
/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 kg
(1
). 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 kg
(1
.
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 Á
/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 AsÁ
/PÁ/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 kg
(1
in
dry matter on soil spiked with 1500 mg As kg
(1
(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
Effect Species Response Experimental set up
Hydroponic Pot/column/batch Field
Mobilisation/extractability As
V
Increased 1, 2, 3, 6, 13, 18
In soil As
III
Increased 18
As-plant uptake As
V
Increased 12, 14, 13, 17 15
Decreased 5, 10, 19
Phytotoxicity
Root elongation As

V
Increased 9, 11, 19
As
III
Slightly increased 9
Yield As
V
Increased 4 3, 4, 6, 8
Decreased 3, 14
No response 8, 14 8
Plant height DSMA
a
Decreased 7
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).
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á
/278264
the first place enhance plant growth and secondly
mobilise exchangeable As resulting in increased
total As uptake.
Most studies on AsÁ
/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 (NH
4
Cl-extractable) As and the Al-
(NH
4
F-extractable), Fe- (NaOH-extractable),
and Ca-bound (H
2
SO
4
-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 modified Chang 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 acetateÁ
/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 As
V
in
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).
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278 265
Turpeinen 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 the Wenzel 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
/As
V
B/
MMAAB/As
III
, while Carbonell-Barachina et al.
(1998) obtained the order of DMAAB/MMAA$/
As
V
B/As
III
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 As
III
,As
V
and DMAA. As
III
was readily oxidised to As
V
, 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
As
V
, 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 As-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 2 provides
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 kg
(1
) 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).
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á
/278266
kg
(1
) 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 soilÁ
/plant and rootÁ/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
kg
(1
(Ma et al., 2001). Based on the accumu-
latorÁ
/excluder concept of Baker (1981) As toler-
ant plants (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 l
(1
; 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
P. vittata 22 630 1500
c
15 1
7234 303 97 74 23.8 1
755 6 126 1
P. calomelanos 8000 88 135 59 91 2
Tolerant plants (non-accumulators)
A. capillaris 3470 26 500 0.13 3
Agrostis catellana 170 1000 17 000 0.01 0.17 4
Agrostis delicatula 300 1800 17 000 0.018 0.17 4
Cynodon dactylon 1600 10 850 9530 0.17 0.15 5
Paspalum tuberosum 1130 7670 0.147 6
Spergularia grandis 1175 7670 0.15 6
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.
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á
/278 267
seem to be also involved in detoxification of As in
Silene vulgaris (Sneller et al., 1999)andNicotiana
tabacum (Nakazawa et al., 2000). Most As in
fronds of P. vittata and P. calomelanos is present
as As
III
, whereas As
V
dominates in roots (Ma et
al., 2001; Francesconi et al., 2002). As metabolism
and complexation in plants havebeenreviewed 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 symbiotic 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 As
V
,As
III
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 As
III
. Twenty-four-hours uptake of As
V
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
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278268
knowledge it can be concluded that mycorrhizal
fungi in most cases tend to limit pollutant transfer
to the host plant, while improving plant nutrition
(Leyval et al., 1997). However, detailed investiga-
tions are needed to address this issue for mycor-
rhizal hyperaccumulators.

Numerous bacteria, fungi, yeasts and algae are
able to transform As compounds by oxidation,
reduction, demethylation and methylation (Fran-
kenberger and Arshad, 2002). Microbial reduction
of As
V
to As
III
is known to occur by dissimilatory
reduction and detoxification activities of microbes.
In the process of dissimilatory reduction As
V
is
utilised as a terminal electron acceptor during
anareobic respiration (Dowdle et al., 1996). De-
toxification of As by microbes involves an As
V
reductase and As
III
extrusion by an As
III
-efflux
pump (Cervantes et al., 1994). Inskeep et al. (2002)
suggested that reduction of As
V
to As
III
via the
detoxification pathway may be a widely-distribu-
ted trait of soil and aquatic microbes. However,

the predominant As-species found in soil solution
of aerobic soil remains As
V
(Wenzel et al., 2002a).
Bacterial oxidation has been described for many
species and could be a feasible remediation strat-
egy as As
V
is less toxic and less mobile than As
III
(Frankenberger and Losi, 1995). Microbial methy-
lation is common for both bacteria and fungi and
has been known for a long time (Cheng and Focht,
1979; Cullen et al., 1984). Formation of volatile
methylarsines offer the possibility to employ As
volatilisation as remediation strategy. However,
only marginal losses by formation of gaseous
arsines have been reported for a Na cacodylate
and methanarsonic-acid-amended soil (Goa and
Burau, 1997), for a soil polluted by As-containing
wood preservatives (Turpeinen et al., 1999) and a
soil containing geogenic As (Prohaska et al., 1999).
The number of microorganisms in the rhizo-
sphere is typically one order of magnitude larger
than in non-rhizosphere soil due the continuous
input of root-derived organic substrate, resulting
in a more diverse, active and synergistic commu-
nity (Marschner, 1995). Root-microbe-induced
transformation processes may therefore affect the
fate of As in the rhizosphere which has not yet

been explored.
2.9. Precipitation phenomena in the rhizosphere
Water flow, ion transport by convection and
diffusion processes, plant uptake, changes of pH
and redox potential, root exudation, etc. alter the
chemical composition of the soilÁ
/root interface
and may result in precipitation phenomena, fa-
vouring pollutant immobilisation. Formation of
Fe plaque in the rhiszophere of salt marsh plants
caused by leakage of O
2
via aerenchyma resulted
in a effective fixation and detoxification of As
(Doyle and Otte, 1997). Precipitation of pyromor-
phyte, Pb
5
(PO
4
)Cl, has been reported in both the
rhizsophere and outer cell walls of the root
epidermis of A. capillaris (Cotter-Howells and
Caporn, 1996; Cotter-Howells et al., 1999). No
formation of defined As mineral phases in the
rhizosphere has been reported yet.
2.10. Conceptual model of the fate of As in the
soilÁ
/rhizosphereÁ/plant system
Based on the review of available literature and
theoretical considerations presented in the pre-

vious sections, we propose a conceptual model of
the fate of As in the soil Á
/rhizosphereÁ/plant
system including As uptake, chemical speciation
in soil solution and interactions with the soil solid
phase and plant nutrition of P and Fe (Fig. 2). The
Fig. 2. Conceptual model of As in the soil Á/rhizosphereÁ/plant
system.
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á
/278 269
individual processes depicted in the model as
fluxes/transformations (solid lines) and influ-
ences/interactions (dotted lines) are well estab-
lished, but yet have not been applied to describe
the fate of As in the soil Á
/rhizosphereÁ/plant
system. This model highlights key processes and
their interactions and may provide some guidance
in future research on the fate of As in the rhizo-
sphere of terrestrial plants.
Arsenic in soils is largely associated with Fe
oxides/hydroxides. The pH (H
'
) in the rhizo-
sphere may differ up to two units from that in bulk
soil. Under aerobic conditions As is mainly present
as As
V
in soil solution which is desorbed from
adsorption sites upon pH increase. Both, plant-

induced reductions of the redox potential (e
(
)and
drastic pH decreases in the rhizosphere may
dissolve Fe oxides/hydroxides, resulting in con-
comitant release of As and P into the soil solution.
Redox potential and pH control the prevailing
redox and hydrolysis species of As in solution.
Reduction of As
V
to As
III
would result in en-
hanced bioavailability and toxicity to plants (Car-
bonell-Barachina et al., 1999; Marin et al., 1993).
Fe and As uptake may interact because root-
induced mobilisation processes that increase Fe
solubility from Fe oxide/hydroxide phases inevi-
tably result in co-dissolution of Fe-bound As. Two
different types of root response, namely strategy I
and II, have been described as mechanisms for Fe
acquisition (Marschner and Ro
¨
mheld, 1994). Par-
ticularly the release of reducing and chelating
compounds, a mechanisms of strategy I plants,
would increase As bioavailability in the rhizo-
sphere.
Carboxylic acids (RÃ
/COOH) released by P-

deficient plants have been reported to be involved
in mobilisation of inorganic P in the rhizosphere
(Hoffland, 1992; Neumann and Ro
¨
mheld, 1999;
Kirk et al., 1999; Dinkelaker et al., 1995). Such
processes are also likely to affect As availability
due to the well known physicochemical similarities
between arsenate and phosphate resulting in
competition for sorption sites in soil (Adriano,
2001) and absorption via the P-specific uptake
system (Asher and Reay, 1979; Meharg and
Macnair, 1990).
2.11. Critique of the use of solution culture for
studies on As uptake by plants
Many investigations on plant and mycorrhizal
response to As are based on hydroponic experi-
ments. Such experiments can be set up easily and
growth conditions are clearly defined if carried out
in a controlled environment. Researchers have
examined As-induced toxicity, As Á
/P interactions,
nutritional implications of excess As supply,
uptake kinetics, bioavailability of different As
species, genetics of As-tolerance etc. However,
one should note that many investigations put
emphasis on As concentrations that are not
common in real world conditions. Fig. 3 gives an
overview of As concentrations employed in such
solution culture experiments, compared with nor-

mal concentrations in soil solution (Sadiq, 1997;
Wenzel et al., 2002a), the WHO drinking water
standard (WHO, 1994), the highest reported value
of As in field-collected soil solutions in aerobic
soils (Wenzel et al., 2002a) and the largest
concentration found in ground waters in Bangla-
desh (Chakraborti et al., 2001).The smallest con-
centration used, except in control solutions, was
about 20 mgl
(1
(Cox et al., 1996), the largest
375 000 mgl
(1
(Meharg and Macnair, 1990). The
ranges as summarised in Fig. 3 are commonly
found in literature (Meharg and Macnair, 1994).
There is virtually no literature available on field-
collected As soil solutions, except Wenzel et al.
(1997a,b, 2002a). This is remarkable as differences
in concentrations can be found between lysimeter-
collected soil solutions and water saturation ex-
tracts, the latter being substantially larger (Wenzel
et al., 1997b). Generally, total concentrations of
As in soil solutions are usually in the range of 1Á
/4
mgl
(1
even in moderately contaminated soils
(Sadiq, 1997; Wenzel et al., 2002a). Concentra-
tions used in hydroponic experiments by far

exceeded the highest concentration found in
field-collected soil solutions. Wenzel et al.
(2002a) reported maximum concentrations of 147
and 171 mgl
(1
in the ABw and Cw horizon of a
Calcaric Cambisol (total soil As 2250 mg kg
(1
,
geogenic contamination). The largest concentra-
tion of 4730 mgAsl
(1
in ground waters of the
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278270
Ganges delta was measured in Bangladesh (Chak-
raborti et al., 2001).
For future research it is necessary to include
concentrations in experiments typically found in
natural soil. This is less critical for genetic studies
on metal/metalloid tolerance (e.g. Macnair et al.,
1992), but is certainly required if the fate of As in
the rhizosphere is targeted by such experiments.
The studies of Meharg et al. (1994), Wright et al.
(2000) show that conclusions drawn from solution
culture experiment may contradict those found in
experiments using soil-grown and mycorrhizal
plants.
3. Potential application of arsenic transformation
processes for soil remediation purposes
3.1. Phytoextraction

Phytoextraction is the use of pollutant-accumu-
lating plants which are able to extract and
translocate pollutants to the harvestable parts.
Phytoextraction can be divided in continuous
phytoextraction (using hyperaccumulator plants)
and induced phytoextraction (chemically induced
accumulation of metals to crop plants). Induced
phytoextraction has not yet been applied to As.
This technique potentially threatens deeper soil
layers and ground waters by the artificially in-
duced mobilisation of pollutants (Sun et al., 2001;
Wenzel et al., 2002b) and is therefore not con-
sidered as a reasonable option for remediation of
As-contaminated soils.
The most abundant group of hyperaccumula-
tors, those that hyperaccumulate Ni, comprise
about 318 taxa (Baker et al., 2000; Brooks,
1998). In contrast, the first As-hyperaccumulating
plants, the ferns P. vittata and P. calomelanos,
have been discovered only recently. Both plants
produce large biomass and are therefore promising
candidates for phytoextraction purposes (Ma et
al., 2001; Francesconi et al., 2002; Visoottiviseth et
al., 2002). No replanting after harvest would be
required as both ferns are perennial, which makes
it even more cost-effective. Similarly to agricul-
tural practices, cultivation of As-hyperaccumula-
tors has to be optimised in order to provide a
efficient and cost effective alternative to common
engineering-based remediation technologies. This

includes breeding, developing management prac-
tises of perennial ferns and rhizosphere manipula-
tion. The natural occurrence of the newly
discovered As hyperaccumulators is mainly re-
Fig. 3. Range of As concentrations commonly used in solution culture experiments.
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á
/278 271
stricted to tropical and subtropical regions of the
world. However, P. vittata is also native to Spain,
Italy and Greece (Jalas and Suominen, 1972)and
has been naturalised on a hot colliery tip in
England since before 1924 (Stace, 1991).
3.2. Phytostabilisation
Phytostabilisation refers to the use of pollutant-
tolerant plants for mechanical stabilisation of
polluted land in order to prevent bulk erosion,
reduce air-borne transport and leaching of pollu-
tants. As-tolerant plants that may be potentially
used for phytostabilisation purposes havebeen
known for a long time (Porter and Peterson, 1975;
Rocovich and West, 1975; Benson et al., 1981).
Depending on the origin of As pollution,
elevated As concentrations often coincide with
pollution by other contaminants such as Cu or Zn.
This implies that plants used for phytostabilisation
of contaminated land need to exhibit multiple
tolerance to metals and metalloids. Examples of
such species which haveevolved multiple tolerance
to various metals and As include A. capillaris
(Porter and Peterson, 1975; Benson et al., 1981;

Symeonidis et al., 1985), D. cespitosa (Cox and
Hutchinson, 1980, 1981) and S. vulgaris (Paliouris
and Hutchinson, 1991). In contrast to phytoex-
traction, plants are required that take up only
small amounts of As and other metals in order to
prevent transfer into the wild-life food chain.
Sharples et al. (2000a) suggested that an As-
tolerant mycorrhizal fungus may restrict As trans-
fer to the host plant, while enhancing plant P
status. Specific inoculates of As-tolerant mycor-
rhiza may facilitate revegetation of sites severely
polluted with As.
3.3. Phytoimmobilisation
Phytoimmoblisation is the use of plants to
decrease the mobility and bioavailability of pollu-
tants by altering soil factors that lower pollutant
mobility by formation of precipitates and insoluble
compounds and by sorption on roots. Based on
the chemical similarities of As and P, there may be
precipitate formation of As Á
/Pb compounds as
shown for PÁ/Pb precipitates in the rhizosphere of
A. capillaris (Cotter-Howells and Caporn, 1996;
Cotter-Howells et al., 1999). Other plant-mediated
processes of As immobilisation at the soilÁ
/root
interface are pH decrease and oxidation of the root
environment by O
2
release from roots. Doyle and

Otte (1997) found accumulation of As on Fe
plaque in the oxidised rhizosphere of salt marsh
plants which may provide an effective immobilisa-
tion and detoxification mechanism.
3.4. Phytovolatilisation
Phytovolatilisation is the use of plants to
volatilise pollutants and has been demonstrated
for Hg and Se. In case of Hg this was achieved by
genetic manipulation of plants (Rugh et al., 1996)
whereas phytovolatilisation of Se occurs naturally
in plants (Terry and Zayed, 1994). De Souza et al.
(1999) demonstrated that rhizosphere bacteria can
enhance Se volatilisation and accumulation in
plants. Volatilisation of As is also known to occur
in natural environments (Frankenberger and Ar-
shad, 2002), but rhizosphere studies are still
missing to assess if formation of gaseous arsenicals
is enhanced at the soilÁ
/root interface. Available
information on As volatilisation for soil suggests
that in the absence of plant roots volatile com-
pounds account only for small proportions (B
/1%)
of total As (Goa and Burau, 1997; Prohaska et al.,
1999; Turpeinen et al., 1999). However, it has to be
considered that methylated Se is nontoxic, whereas
volatile arsines are still toxic (Frankenberger and
Arshad, 2002)
4. Conclusions and research needs
Arsenic represents a pollutant of major concern

throughout the world. Rhizosphere processes are
deemed to play a key role in the improvement of
phytoremediation technologies. Substantial pro-
gress has been achieved during the last decades to
understand the fate of As in the soil Á
/plant system.
However, virtually no literature is available that
particularly addresses transformation processes
and interactions of As in the rhizosphere.
Future investigations should:
W.J. Fitz, W.W. Wenzel / Journal of Biotechnology 99 (2002) 259 Á/278272
. be based on rhizobox-type experiments (e.g. Li
et al., 1991; Wenzel et al., 2001b) rather than
solution culture studies;
. cover realistic As concentrations, found in
natural soil solutions;
. put emphasise on As, P, and Fe interactions in
the rhizosphere;
. examine rates and quality of root exudates
potentially involved in As hyperaccumulation
and/or As tolerance;
. investigate the role of mycorrhizas in As
tolerance and Á
/hyperaccumulation;
. investigate the role of soil microbes on As
V
/
As
III
transformations in the rhizosphere;

. develop rhizosphere management technologies
(e.g. based on knowledge obtained in rhizobox
studies) in order to facilitate and improve
practical applications of phytoremediation of
As-polluted soils.
The accidental discovery of As-hyperaccumulat-
ing ferns may have been surprising for many
researchers, suggesting that systematic screening
may reveal other plants that hyperaccumulate As.
Given the unexplored potential, emphasis should
be put to screening programs rather than on
genetical engineering of the few known As hyper-
accumulators.
Acknowledgements
This study was supported by the University of
Agricultural Sciences Vienna, Priority Research
Project No. 16.
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