The Responses of Plants to Metal Toxicity

The Responses of Plants to Metal Toxicity:
A review focusing on Copper, Manganese and
Zinc
S. M. Reichman
Published by the
AUSTRALIAN MINERALS & ENERGY ENVIRONMENT FOUNDATION
1
Published in Melbourne, Australia by:
Australian Minerals & Energy Environment Foundation
144 High St Prahran VIC 3181
Melbourne, Victoria, 3000
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Published as Occasional Paper No.14
ISBN 1-876205-13-X
© The Australian Minerals & Energy Environment Foundation 2002
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Enquiries should be directed to the Australian Minerals & Energy Environment Foundation.
2
The Responses of Plants to Metal Toxicity.
A Review Focusing on Copper, Manganese and Zinc
Suzie Reichman
Table of contents
1.
2.
3.
Abstract ........................................................................................................................................................... 5
Introduction..................................................................................................................................................... 7
Soil and metals ................................................................................................................................................ 8
3.1
Factors affecting metal availability ......................................................................................................... 8
3.1.1
Total metal concentration................................................................................................................ 8
3.1.2
pH.................................................................................................................................................... 8
3.1.3
Organic matter................................................................................................................................. 9
3.1.4
Clays and hydrous oxides................................................................................................................ 9
3.1.5
Oxidation and reduction .................................................................................................................. 9
3.2
Soil testing............................................................................................................................................. 10
3.2.1
Measures of total metal ................................................................................................................. 11
3.2.2
Chelation techniques ..................................................................................................................... 11
3.2.3
Ion exchange resins ....................................................................................................................... 12
3.2.4
Neutral salt extractants .................................................................................................................. 12
3.2.5
Soil solution and pseudo-measures of soil solution....................................................................... 13
3.2.6
Speciation and toxic metal forms .................................................................................................. 14
3.2.7
Plant factors................................................................................................................................... 14
3.2.8
Conclusions on the bioavailability of metals in soils .................................................................... 14
3.3
Relationship between soil and solution culture studies ......................................................................... 15
3.3.1
Solution culture as a model of the soil (solution).......................................................................... 15
3.3.2
Comparing toxicity in soil with that in solution culture................................................................ 15
4. Plants and metals........................................................................................................................................... 16
4.1
Plant uptake and transport of metals ..................................................................................................... 16
4.1.1
Uptake mechanisms ...................................................................................................................... 16
4.1.2
Transport within the plant ............................................................................................................. 18
4.2
Relationship between metal in the growth substrate and growth .......................................................... 20
4.3
Relationship between tissue concentration and growth......................................................................... 20
4.4
Symptoms and visual evidence of toxicity............................................................................................ 22
4.4.1
Copper........................................................................................................................................... 22
4.4.2
Manganese .................................................................................................................................... 23
4.4.3
Zinc ............................................................................................................................................... 23
4.5
Effects of toxicity on physiology .......................................................................................................... 24
4.5.1
Photosynthesis............................................................................................................................... 24
4.5.2
Transpiration and water budgets ................................................................................................... 25
4.5.3
Enzymes and cell metabolism ....................................................................................................... 25
4.5.4
Cell functioning............................................................................................................................. 26
4.6
Mechanisms of tolerance....................................................................................................................... 26
4.6.1
Restriction of uptake or transport.................................................................................................. 26
4.6.2
Compartmentation and complexing within the cell....................................................................... 28
4.6.3
Conclusions on metal tolerance..................................................................................................... 32
5. Methodology ................................................................................................................................................. 33
5.1
Solution culture ..................................................................................................................................... 33
5.1.1
Traditional non-renewed solution culture and dilute, renewed solution culture ........................... 33
5.1.2
pH.................................................................................................................................................. 33
5.1.3
Ionic strength................................................................................................................................. 34
5.1.4
Phosphorus .................................................................................................................................... 34
5.1.5
Iron and chelates ........................................................................................................................... 35
6. Metal toxicity and Australian native plants................................................................................................... 37
6.1
Information on the tolerances and responses of Australian plants to excess Cu, Mn and Zn................ 37
6.2
Within species genetic variability and the significance for metal tolerance.......................................... 37
7. Final comments ............................................................................................................................................. 39
8. References..................................................................................................................................................... 40
3
The Responses of Plants to Metal Toxicity
A Review Focusing on Copper, Manganese & Zinc
Author: S. M. Reichman
PhD Scholar, Centre for Mined Land Rehabilitation, The University of Queensland
Winner of the 2000 Ameef Environmental Excellence Awards–
the Literature Review Award
Metal toxicity issues are of significant concern in many industries, including mining. A literature review was
undertaken to establish the state of knowledge of the responses of plants to metal toxicities. Issues covered
include metal bioavailability and testing, the effects of substrate metal on plant growth, physiology, symptoms
and tissue concentrations, mechanism of tolerance and the use of solution culture as a model of the soil.
4
1. Abstract
Metal contamination issues are becoming increasingly common in Australia and elsewhere, with many
documented cases of metal toxicity including in mining and agriculture. Metals are a natural part of terrestrial
systems occurring in soil, rock, air, water and organisms. A few metals, including Cu, Mn and Zn, are however
essential to plant metabolism in trace amounts. It is only when metals are present in bioavailable forms at
excessive levels that they have the potential to become toxic to plants. If decisions are to be made about how to
utilise potentially toxic substrates, an understanding is required of plant responses, plant sensitivities and
tolerances, and the substrate factors affecting plant responses.
Metals are frequently measured, and regulatory decisions made for toxicity on the basis of the total metal
concentration in a growth substrate. However, when the interest is the toxicity of a metal to plants then the
primary concern must be the bioavailable soil metal fraction. Many factors affect the bioavailability of metals in
soil, the most important being the total metal concentration, pH, the presence of organic matter, redox
conditions, and the presence of clays and hydrous oxides.
Plant roots directly remove nutrients and metals from the soil solution. Attempts to find suitable soil tests for
metal toxicity have varied from measuring concentrations of the immediately bioavailable metal to also
including measures of the capacity of the soil to replenish the metal or buffer the metal concentration as it is
removed from the soil solution. Measures of chelate extractable metals, such as DTPA and EDTA extracted
metal, do not necessarily correlate well with bioavailable metal. Other techniques show promise such as ion
exchange resins, CaCl2-extractable metal, soil solution and free metal ion activity in the soil solution. However,
research evaluating these techniques has tended to be of poor quality and with considerable variation in
methodology within a test type. Standardisation and further research is required to determine the exact nature of
the bioavailable metal and hence develop appropriate soil tests.
Plant responses to metals are dose dependent. For essential metals, these responses cover the phases from
deficiency through to sufficiency/tolerance to toxicity. For non-essential metals, only the tolerance and toxicity
phases occur. The idea of a critical or threshold toxicity is often used to establish the point at which metals
cause significant growth decreases. These are often defined as the metal concentrations corresponding to a yield
decrease of 10%. This principle can be used to determine both the critical substrate concentration and the critical
foliar concentration. Critical concentrations vary considerably across metals and plant species. In plant nutrition
studies, the youngest fully expanded leaf (YFEL) is often used as the standard plant tissue for comparison of
foliar concentrations as trace element concentrations in this tissue are often independent of plant age. However,
very little use has been made of the YFEL, or other standardised tissue, in toxicity research, making comparisons
between studies difficult.
Metal toxicities are accompanied by a range of leaf symptoms which can be used to aid diagnosis. Copper
toxicity often causes foliar interveinal chlorosis, the leaf becoming necrotic with increasing exposure. In Mn
toxicity, symptoms include chlorosis of older leaves, necrotic spotting and a symptom on young foliage known
as “crinkle leaf”. Zinc toxicity symptoms include chlorosis and reddening of younger leaves with necrotic
lesions on leaves in severe cases.
The action of metals is seen at the whole plant level in reduced growth, and at the organ level in leaf symptoms.
At a smaller scale, the effects of metals can be seen as cellular symptoms. Symptoms, both macro and cellular,
and growth effects are side effects of the direct mode of action. The direct mode of action of a metal is on plant
metabolism. Each metal has a different mode of action. However, in general, metal toxicity has been shown to
reduce photosynthesis, affect enzyme and protein production and utilisation, alter nutrient transport and have
negative effects on cellular functioning.
There is much contention in the literature on the possible mechanisms of metal tolerance. This could indicate a
general lack of understanding of metal toxicity issues but is just as likely to reflect the complex nature of higher
plant responses to metal toxicity. It is quite likely that different species may have evolved different mechanisms
to tolerate excess metals, and that even within the one species more than one mechanism could be in operation.
There are a number of strategies that plants could employ to combat high external metal concentrations. These
can be classified into two main categories, i.e. firstly, restriction of uptake or transport and secondly, internal
tolerance mechanisms. Different plant parts, species, and metals appear to elicit different responses and possibly
5
more than one response. Mechanisms appearing to be important include exclusion at the cell-wall, sequestration
in vacuoles, and the production of phytochelatins.
Solution culture is frequently used to determine plant sensitivities to metal toxicity. The main advantage that
solution culture has over soil culture is that the composition of the growing medium can be defined, manipulated
and measured with a high level of precision, thus reducing confounding variables and giving quantitative
answers. The main difference between solution culture and soil culture is that, in solution culture, plant roots are
bathed in a continuously stirred aqueous medium compared to soil where plant roots must continually explore
the soil to gain nutrients and water. To obtain meaningful results in solution culture studies, a number of factors
must be considered. Conditions must be as similar to soil conditions as possible, this includes taking note of the
pH and fertility of comparable soils. As research has shown that precipitated metals are not plant-available, and
that complexed metals are generally not as available as the free metal, the experimental design should be based
on knowledge of the effects of pH, chelators, and solubilities of metals and nutrients. Computer simulation
programs such as GEOCHEM can be used to predict metal and other nutrient solubilities and speciation. Failure
to do this has resulted in many experiments with reduced precision, and hence confidence in results obtained is
limited.
Finally, most research has been undertaken on crop and pasture species with little research into the effects of
metal toxicity on trees. In particular, very little research on metal toxicity has been concerned with Australian
native trees. It is important that we understand the toxicity responses of Australian trees to metals so that we can
utilise appropriate species in the rehabilitation of contaminated areas, identify metal toxicities when they occur,
and regulate metal emissions effectively.
6
2. Introduction
As human activity impacts upon the environment, metal contamination issues are becoming increasingly
common (Fernandes and Henriques, 1991). In Australia, there are many documented cases of metal toxicity,
including in mining (Richards et al., 1996) and agriculture (Grasmanis and Leeper, 1966). Metals are a natural
part of terrestrial systems and occur in soil, rock, air, water, and organisms. A few metals, including Cu, Mn and
Zn, are required by plants in trace amounts. It is only when metals are present in bioavailable forms at excessive
levels that they have the potential to become toxic to plants.
There are many sources of metals in soils and growth media, including :
• Natural e.g. soil parent material, windblown dusts, volcanic eruptions, marine aerosols, and forest fires
(Fergus, 1954),
• Agricultural e.g. fertilisers, sewage sludges and animal wastes used as fertilisers, pesticides and irrigation
water (Grasmanis and Leeper, 1966; Reddy et al., 1995),
• Energy production e.g. emissions from power stations (Gorzelska, 1989),
• Mining and smelting e.g. tailings, smelting and refining, transportation (Brooks et al., 1992; Helmisaari et
al., 1995),
• Secondary metal production and recycling operations e.g. melting of scrap, refining, plating alloying
(Seaward and Richardson, 1990),
• Urban/ industrial complexes e.g. incineration of wastes and waste disposal (Milbocker, 1974; Brooks et al.,
1992),
• Automobiles e.g. combustion of petroleum fuels (Lagerwerff and Specht, 1970; Gorzelska, 1989).
Generally, metal toxicity issues do not arise in natural soils with their native vegetation. Even if the soil is
naturally high in a particular metal, native plants will often have become adapted over time to the locally
elevated levels (Brooks et al., 1992; Ouzounidou et al., 1994). However, if humans bring new growth regimes,
such as agriculture, with plants not evolved on these specialised soils, then toxicity issues can develop (Fergus,
1954). Most metal toxicity occurs as a result of anthropogenic disturbance, such as mining, where unnaturally
high amounts of metals are released during various processes (e.g. Helmisaari et al., 1995). As a result of the
strong influence of pH on metal solubility (McBride et al., 1997), anthropogenic processes which result in the
lowering of substrate pH can cause metal toxicities, even if no extra metal has been added to the system (Fergus,
1954; Kelly et al., 1990; Robinson et al., 1995).
There are documented cases of many different metals causing toxicity issues (e.g. Grasmanis and Leeper, 1966;
Godbold and Huttermann, 1985; Merry et al., 1986b; Kelly et al., 1990). Copper, Mn and Zn are three metals
known to cause metal toxicities both in Australia and around the world. Documented cases of Cu, Mn or Zn
contamination in Australia include those at ex-mining sites (Richards et al., 1996; Roseby et al., 1998;
Lottermoser et al., 1999), in agricultural regions (Grasmanis and Leeper, 1966; Merry et al., 1986a), at
industrial sites (Phillips and Chapple, 1995), in urban areas (Markus and McBratney, 1996), along coasts and
waterways (Norris, 1986; Hanley and Couriel, 1992) and in naturally elevated soils (Lottermoser, 1997).
The majority of research on plant responses to excess metals has been undertaken on crop and pasture species
with little research into the effects of metal toxicity on trees. In particular, very little research on metal toxicity
has studied Australian native trees. Australia has a unique flora, which may react differently to metals than other
trees and especially crop and pasture species. If decisions are to be made about how to utilise potentially metaltoxic substrates, an understanding is required of plant responses, plant sensitivities and tolerances, and the
substrate factors affecting plant responses.
7
3. Soil and metals
3.1 Factors affecting metal availability
Plants cannot usually access the total pool of a metal present in the growth substrate. Instead, that fraction of the
metal which plants can absorb is known as the available or bioavailable fraction. Metals present in a soil can be
divided into a number of fractions including; the soluble metal in the soil solution, metal-precipitates, metal
sorbed to clays, hydrous oxides and organic matter, and metals within the matrix of soil minerals. These
different fractions are all in dynamic equilibrium with each other (Norvell, 1991). However, while the soluble
metal in the soil solution is directly available for plant uptake other soil metal pools are less available (Davis and
Leckie, 1978; del Castilho et al., 1993). For example, change in the concentration of metal in the matrix of soil
minerals is slow relative to exchange and desorption reactions between clays, hydrous oxides, organic matter and
the soil solution (Shuman, 1991; Whitehead, 2000).
Metals within the soil solution are the only soil fraction directly available for plant uptake (Fageria et al., 1991;
Marschner, 1995; Whitehead, 2000). Hence, factors which affect the concentration and speciation of metals in
the soil solution will affect the bioavailability of metals to plants. Soil factors which have an affect on metal
bioavailability include the total metal present in the soil, pH, clay and hydrous oxide content, organic matter and
redox conditions.
3.1.1 Total metal concentration
The total metal concentration of a soil includes all fractions of a metal, from the readily available to the highly
unavailable. Other soil factors, such as pH, organic matter, clay and redox conditions, determine the proportion
of total metal which is in the soil solution. Hence, while total metal provides the maximum pool of metal in the
soil, other factors have a greater importance in determining how much of this soil pool will be available to plants
(Wolt, 1994). In addition, researchers have found that while total metal correlates with bioavailable soil pools of
metal it is inadequate by itself to reflect bioavailability (Lexmond, 1980; Sauve et al., 1996; McBride et al.,
1997; Sauve et al., 1997; Peijnenburg et al., 2000)
3.1.2 pH
The equilibrium between metal speciation, solubility, adsorption and exchange on solid phase sites is intimately
connected to solution pH (Olomu et al., 1973; Kalbasi et al., 1978; Cavallaro and McBride, 1984; Sauve et al.,
1997). Hence, numerous studies have found soil pH to have a large effect on metal bioavailability (Turner,
1994; McBride et al., 1997).
Both Mn and Zn bioavailability are strongly affected by soil pH (Fergus, 1954; McGrath et al., 1988; Turner,
1994). As soil pH decreases, Mn and Zn must compete with the extra H+ and Al3+ for positions on the exchange
sites, solubility of Mn and Zn increases in the soil solution and a greater proportion is present as highly available
free metal ions in the soil solution (Kalbasi et al., 1978; McBride, 1982; Bar-Tal et al., 1988; Msaky and
Calvet, 1990; Sauve et al., 1997). This increases the concentrations of Mn and Zn in the directly bioavailable
fraction, i.e., the soil solution (Jeffery and Uren, 1983). In accordance with the changes in metal bioavailability
associated with a change in pH, many studies have found that plant uptake of Mn and Zn increases as soil pH
decreases. Hence, in Zn contaminated soils as pH decreased Zn concentration increased in shoots of Arachis
hypogaea (peanut) (Parker et al., 1990; Davis-Carter and Shuman, 1993) and the potential for Mn toxicity in
Phaseolus vulgaris (bean) (Fergus, 1954) and Vigna unguiculata (cowpea) (Vega et al., 1992) increased in acid
soils.
While solution pH affects Cu speciation, solubility, complexation and adsorption (Payne and Pickering, 1975;
Msaky and Calvet, 1990; Reddy et al., 1995) some soil studies have found little relationship between soil pH
and Cu concentration in the soil solution (Jeffery and Uren, 1983; McGrath et al., 1988; Sauve et al., 1997).
The reason for this is the strong affinity of Cu for organic matter (Norvell, 1991). Therefore, the amount of
organic matter dissolved in the soil solution, especially in soils high in organic matter, can be a more important
determining factor on Cu solubility than pH (see section 3.1.3 for further discussion)
8
3.1.3 Organic matter
-
Metal ions can be complexed by organic matter altering their availability to plants. The COO groups in both
solid and dissolved organic matter form stable complexes with metals (Stevenson, 1976; Baker and Senft,
1995). Hence, as the amount of organic matter present in soil increases the opportunity for forming stable metalorganic matter complexes increases. In general, plants are unable to absorb the large metal-complexes and so the
bioavailability of metals decreases.
Copper ions forms strong coordination complexes with organic matter (Stevenson, 1976; Stevenson, 1991).
Hence, Cu is often predominantly found bound to the organic matter fraction in the soil and soil organic matter
can be the most important soil factor in determining Cu bioavailability (del Castilho et al., 1993). In a
Chernozem, between 37 and 91% of the total soil Cu was present in the organic fraction depending on level of
Cu contamination (Pampura et al., 1993). In a range of Cu contaminated soils greater than 98% of the Cu in the
soil solution was bound to organic complexes, irrespective of pH (Sauve et al., 1997). Also, in a different range
of soils, approximately 95% of soil solution Cu was complexed, irrespective of pH. (Fotovat et al., 1997).
Reddy (1995) found the proportion of Cu bound to organic matter in the soil solution increased from 37 to 95%
as the pH decreased. In addition, Cu applied as sewage sludge was retained in the soil solution in greater
quantities than Cu applied as a sulphate because it was bound to dissolved organics from the sludge (Miller et
al., 1987) and the activity of the highly available Cu2+ has been inversely correlated with soil organic matter
(McBride et al., 1997).
The amount of organic matter found in soils also affects the bioavailability of Zn (Shuman, 1975; Bar-Tal et al.,
1988; del Castilho et al., 1993). However, while Zn readily forms complexes with organic matter it does not
compete for these sites as well as Cu (Cavallaro and McBride, 1984) and other more prevalent cations such as
Ca2+ (Fotovat et al., 1997). In the same Chernozem experiment as described in the paragraph above, between 2
and 6% of the total soil Zn was found in the organic fraction (c.f. 37 to 91 % of Cu) (Pampura et al., 1993). In
soil solution, the activity of the highly bioavailable Zn2+ in the soil solution decreased as organic matter
increased across a range of contaminated soils (McBride et al., 1997). Across a range of soils greater than 50%
of the soil solution Zn was present as the free ion (Lorenz et al., 1997). Also, soil solution Zn was found to be
between 23 and 93% organically complexed dependent on soil pH (Reddy et al., 1995). Hence, organic matter,
while important, does not tend to be as big a factor as pH in determining Zn bioavailability (Elrashidi and
O'Connor, 1982).
Manganese tends to form weak coordination complexes with organic matter (Olomu et al., 1973; McBride,
1982). This means that Mn2+ is unable to compete effectively with Cu2+, Zn2+ and other more prevalent cations,
such as Ca2+ and Mg2+ for sites on organic matter and, hence, less Mn is generally found bound to organic matter
than for Cu and Zn (McGrath et al., 1988). For example, across a range of soils approximately 30% of the soil
solution Mn was present as organic complexes (Olomu et al., 1973). Complexed Mn in the soil solution of a
sandy loam increased from 10 to 55% as the amount of organic matter in the soil increased (McGrath et al.,
1988). In contrast, only one study has ever found Mn-organic matter complexation to approach that of Cu, i.e.
Geering et al. (1969) found between 84 and 99% of the Mn present in the soil solution to be bound to organic
matter in soils from a variety of areas. In the majority of cases, organic matter has less importance in the
bioavailability of Mn than for Zn and especially, Cu.
3.1.4 Clays and hydrous oxides
Clays and hydrous oxides, i.e. oxides of Al, Fe and Mn, play an important role in the availability of metals.
Clays and hydrous oxides determine metal availability mainly by specific adsorption to surface hydroxyl groups
(Miller et al., 1987; Pampura et al., 1993), nonspecific adsorption (exchange) (Kalbasi et al., 1978; Basta and
Tabatabai, 1992), coprecipitation (Martinez and McBride, 1998), and precipitation as the discrete metal oxide or
hydroxide (Martinez and McBride, 1998). Hence, increasing clay and hydrous oxide contents in soils provides
more sites for adsorption of metals thus reducing the directly bioavailable metal (Shuman, 1975; Ghanem and
Mikkelsen, 1988; Barrow, 1993; Qiao and Ho, 1996).
3.1.5 Oxidation and reduction
The oxidation/reduction (redox) conditions of a soil can play a role in the availability of metals. The redox
status of the soil can be affected by many factors including waterlogging and compaction. Redox conditions can
affect the availability of metals by affecting the proportion of particular metal species (e.g. Mn(II) vs. Mn(IV) in
the soil solution and by affecting the solubility of metals in the soil solution (Patrick and Jugsujinda, 1992;
Evangelou, 1998).
9
Redox state has a large affect on Mn speciation and solubility in the soil solution (Geering et al., 1969; Olomu
et al., 1973; Sajwan and Lindsay, 1986). Manganese can exist in soil as Mn(II), Mn(III) and Mn(IV), however
only the reduced Mn(II) form is stable in solution (Lindsay, 1979; Evangelou, 1998; Whitehead, 2000).
Manganese (II) is the most soluble form of Mn and so under reducing conditions higher concentrations of Mn2+
will be present in the soil solution (Patrick and Jugsujinda, 1992). Conversely, under more oxidising conditions,
soil solution concentrations of Mn decrease because the equilibrium shifts in favour of Mn(III) and Mn(IV)
which tend to exist mainly as insoluble hydroxides and oxides. For example, increasingly reduced conditions
corresponded with an increase in the highly bioavailable Mn2+ in the soil solution and a corresponding increase
in Mn uptake by Oryza sativa (rice) plants (Schwab and Lindsay, 1983). Under waterlogged conditions
increases in Mn uptake and symptoms of Mn toxicity have been noted in Malus sp. (apple) and Pyrus sp. (pear)
trees (Grasmanis and Leeper, 1966). Hence, reducing soil conditions, such as flooding and soil compaction, tend
to increase the availability of soil Mn and enhance toxicity (Cheng and Ouellette, 1971).
Most Cu and Zn are present as the divalent form in soils with the monovalent forms being highly unstable
(Knezek and Ellis, 1980; Whitehead, 2000). Hence, neither Cu nor Zn tend to be significantly reduced under
low redox conditions (Moraghan and Macagni Jnr, 1991; Whitehead, 2000). However, Zn deficiency has been
noted in flooded soils and it was suggested that this was due to the precipitation of Zn as compounds such as
ZnFe2O4 under reducing conditions (Sajwan and Lindsay, 1986; Moraghan and Macagni Jnr, 1991). As such,
redox conditions play a smaller role in the availability of Cu and Zn when compared with Mn.
3.2 Soil testing
The immediate source of nutrient and metal ions to a plant is from the soil solution (Fageria et al., 1991;
Marschner, 1995; Whitehead, 2000). The percentage of any metal occurring in the soil solution is usually small
compared to the total metal pool in the soil (McGrath et al., 1988; Pampura et al., 1993; Sauve et al., 1997).
Metals in the soil solution are in dynamic equilibrium with the larger soil solid fraction, and so, as metals are
removed from the soil solution by plant uptake, or by other processes such as leaching, replenishment of the soil
solution will occur. This replenishment can come from exchangeable ions, adsorbed salts, precipitated
compounds, mineralisation of organic matter and weathering of soil minerals (Pearson, 1971; Whitehead, 2000).
To adequately assess the toxicity of a metal to plants, analytical approaches must quantify that portion of the
total pool of metal in the soil which the plant can access, i.e. the bioavailable fraction. The immediately
bioavailable soil fraction is within the soil solution. However, because the soil is in dynamic equilibrium
measures of the ability of the soil to buffer and/or to replenish the concentration of metals in the soil solution
could be important in determining bioavailability over a longer time frame. The criterion to assess whether
metals present in soils are at toxic levels is not standardised world-wide and varies from one country and land
use to another (Ross and Kaye, 1994). Beckett (1989) reviewed methodologies utilised to measure metals and
listed tests under 23 categories, with many categories containing numerous techniques. Even this was not an
exhaustive list since techniques such as ion exchange resins were omitted. Most soil tests fall into one of the
following four categories, listed in order from largest to smallest fraction of soil metal measured:
1. Measures which incorporate the total metal,
2. Measures which include the immediately bioavailable fraction plus the long term ability of the soil to
replenish the immediately bioavailable fractions,
3. Measures which include the immediately bioavailable fraction plus the ability of the soil to buffer the
immediately bioavailable fraction as it is utilised, and
4. Measures which include the immediately bioavailable fraction.
It is outside the scope of this review to analyse the applicability of every technique for measuring bioavailable
metal, and so, either commonly used methods or methods which show promise for wider use have been chosen
to represent each category. Hence, this review will focus on the theory behind, and applicability of, measuring
(a) total metal (category 1), (b) chelation techniques (category 2), (c) ion exchange resins (category 3), (d)
neutral salts (category 3), (e) soil solution metal (category 4) and (f) the free metal activity in the soil solution
(category 4).
In general, for a soil test to be successful at measuring the bioavailable soil fraction it must fulfill the following
criteria:
1. Correlate well with plant responses (e.g. tissue concentration or yield parameters) across a wide range of
plant species,
2. Correlate well across a wide range of soil types, and
3. Give consistently good results.
The above factors have been considered when evaluating the following soil tests.
10
3.2.1 Measures of total metal
Metals are frequently measured, and regulatory decisions made, for toxicity, on the basis of the total metal
concentration in a growth substrate (Gupta et al., 1996). Different extractants, generally “strong acids” such as
HNO3, HF, HClO4 and aqua regia, have been utilised to determine the total or “pseudo-total” metal in a soil
(Becket, 1989). However, as discussed in section 3.1, many factors control the availability of metals to plants,
making total metal a poor measure of bioavailable metal. It is incorrect to assume that total metal concentration
is just an indirect measure of available metal (Miles and Parker, 1979; Pampura et al., 1993).
Little correlation (r2=0. 23) was found between total soil Cu and the dry weight of Fallopia convolvus (black
bindweed) in a soil contaminated with a range of Cu concentrations (Kjaer et al., 1998). Plant metal
concentration did not correlate well with the total soil concentration of Cu (r2=0.22), Zn (r2=0.31), Cd (r2=0.23)
or Ni (r2=0.05) in 46 soils with varying degrees of contamination (Haq et al., 1980). In one study on A.
hypogaea, a poor correlation was found between shoot Zn and total soil Zn (r2=0.08) (Davis-Carter and Shuman,
1993). In contrast, relatively good correlations were found in another study between total soil Zn and A.
hypogaea leaf Zn, i.e. r2=0.52 for August and r2= 0.65 for September (Parker et al., 1990). However, that two
different relationships were needed for samplings one month apart suggests a lack of robustness in total soil Zn
as a measure of plant available Zn in A. hypogaea.
Some studies have found good relationships between total metal and plant responses (Aery and Jagetiya, 1997).
However these results have tended to be soil specific and, therefore, do not demonstrate the effectiveness of total
metal as a tool for measuring the effects of excess metal on plants.
3.2.2 Chelation techniques
The most frequently used methods to determine the soil solution plus the ability of the soil to replenish metals
are those using chelates such as ethylenediaminetetraacetic acid (EDTA) and diethylenetriaminepentaacetic acid
(DTPA) (Lindsay and Norvell, 1978). These tests are good at predicting plant deficiency of metals (Lindsay and
Norvell, 1978) but they do not seem to correlate well with metal toxicity in plants (Miles and Parker, 1979;
Jarvis and Whitehead, 1981; Merry et al., 1986b). While strong correlations between plant growth and chelator
extractant tests have been reported at toxicity levels, the relationship obtained was generally only applicable to
one soil type (Miles and Parker, 1979; Plenderleith and Bell, 1990).
There are a number of reasons for the inability of chelating extractants to correlate well with plant toxicity. At
toxic levels the metals may exceed the chelating ability of the extractant (Bell, 1986). This renders these
methods insensitive to increases in metal. Chelating extractant tests were developed for soils with low organic
matter (Lindsay and Norvell, 1978) and have not been correlated for soils high in organics. Hence, the use of
such tests in sewage treated soils is questionable. The chelator extraction methodology requires buffering at near
neutral pHs (Lindsay and Norvell, 1978; Clayton and Tiller, 1979). Such buffering would change metal-soil
interactions as a result of the marked effect of pH on metal solubility (see section 3.1.2). Extraction with DTPA
has been shown to affect Cu solubility differently at differing pHs dependent on the amount of DTPA relative to
Cu (Vulava et al., 1997). At high additions of Cu the solubility was barely affected by pH (Vulava et al., 1997).
In metal contaminated situations DTPA extracted Cu and Pb have been shown to increase with increasing pH
while plant uptake decreased (Merry et al., 1986b). Thus the benefits of DTPA extraction as a tool to measure
metals at toxic concentrations are doubtful, especially in metal addition experiments.
Pederson et al. (1980) developed an EDTA extraction methodology at pH 4.6. However they did not correlate
their method against plant responses, hence the ability of the test to measure plant available metal can not be
evaluated. The EDTA soil test was adapted for use in toxic soils by Clayton (1979). However, results were not
correlated with plant responses and only one Cu-contaminated soil was used. In the one toxic Cu sample tested,
EDTA was shown to extract 76% of the total Cu while in other contaminated soils EDTA was found to extract
between 24-98% of total Zn, 74-93% of Pb and 63-100% of Cd. These amounts suggest removal of metals from
strongly adsorbed and less-labile soil pools which are unlikely to contribute in a significant way to the
replenishment of soil solution metal and therefore are not readily bioavailable (Sims, 1986; Pampura et al.,
1993). Recently, Esnaola et al. (2000) attempted to adapt the DTPA test for metal contaminated soils. However,
they did not correlate results with plant responses and so, it is not possible to evaluate the ability of this test to
measure bioavailable metal.
Soil solution plus soil solution replenishment factor tests do not appear to be a good measure bioavailable metal
at toxic levels. It could be suggested that the replenishment of the soil solution from other metal pools is more
likely to have a significant affect on metal availability in deficiency situations. Here, plants absorb a larger
11
proportion of the soil solution metal and so proportionately more metal is released into the soil solution to
maintain equilibrium. Another reason may involve plant produced phytosiderophores which aid plants in
extracting nutrients from larger soil pools at deficiency levels (Marschner, 1995).
3.2.3 Ion exchange resins
Ion exchange resins accumulate ions through soil exchange mechanisms, thus modeling the action of plant roots.
The idea being that a capsule of cation/anion exchange resin placed in the soil would accumulate ions from soil
processes in a similar manner to a living plant root (Yang et al., 1991). Research with ion exchange resins is
showing great promise for use at the deficiency/sufficiency end of the spectrum (Hamilton and Westermann,
1991; Raij, 1998). Recent research with the use of ion exchange resins is also suggesting that this technique
may provide a measure of the ability of the soil solution to buffer metal concentration in toxic situations (Lee et
al., 1996b; Liang and Schoenau, 1996).
While the technique is theoretically good and shows considerable promise there are a number of issues which
need to be addressed before the applicability of ion exchange resin as a measure of bioavailable metal can be
determined. At present inadequate testing of the technique restricts its wider application. Most of research has
been conducted at deficiency/sufficiency levels, or with theoretical chemistry at elevated metal concentrations
(Harper et al., 1998; Hooda et al., 1999). In soils with elevated metal concentrations there are few studies
correlating ion exchange measurements over a number of soils, comparing the ion exchange resin technique with
other leading measures of metal bioavailability, e.g. soil solution metal, and, most importantly, against plant
responses.
One study compared an ion exchange resin technique with HCl, CaCl2 and DTPA extraction for Cu, Cd and Pb
and Triticum aestivum (wheat) seedlings grown in nine soils (Lee and Zheng, 1994). Triticum aestivum uptake
of Cu was most highly correlated with the exchange resin extracted Cu while the highest correlation for Cd was
with CaCl2 and Pb with DTPA. However, while four of the soils were considered ‘contaminated’ by the authors,
they did not result in yield decreases.
Many variations exist within the basic methodology of ion exchange resin extraction. Variations exist in the
method for saturating the resin, in soil to resin ratio, length of shaking time, type of resin, separating the soil
from the resin, extracting the ions from the resin, and in whether the test is made in situ or in a laboratory (Raij,
1998). Standardisation of the technique and finding the methodology most correlated with bioavailable metal
are important issues as, for example, lengthy extraction periods have the potential to remove metals from
strongly adsorbed and less-labile soil pools that are not readily bioavailable (Hamilton and Westermann, 1991;
Agbenin et al., 1999). Therefore, the potential success of ion exchange resins might depend upon how accurately
the methodology can be linked to bioavailability, rather than the inherent applicability of the product itself. At
present there is a lack of research evaluating the relative applicability of the different resin methodologies.
A new ion exchange resin methodology known as diffusive gradients in thin films (DGT) has shown promise
when tested theoretically (Harper et al., 1998; Hooda et al., 1999). The DGT technique involves covering a
layer of chelex exchange resin with a diffusive gel layer and a filter layer, thus creating a concentration gradient
in the diffusion layer and theoretically modelling the action of plant roots on the soil solution (Harper et al.,
1998). It remains to be seen how well the technique correlates with plant responses.
If an ion exchange resin exchange technique can be identified which measures bioavailable metal then the
technique has great promise. Ion exchange resin methodologies are easy and inexpensive (Hamilton and
Westermann, 1991) making them a potentially practical solution to the issue of measuring bioavailable metal.
3.2.4 Neutral salt extractants
The use of neutral salts was first utilised on soils at deficiency/ sufficiency concentrations and has been
suggested as a universal extractant (McLaughlin et al., 2000). Numerous neutral salt extractants have been
utilised across the range from deficiency through sufficiency to toxicity, including NaNO3, NH4NO3, NH4Cl,
MgCl2, KNO3 and Ca(NO3)2 with CaCl2 being frequently claimed in the literature to measure bioavailable metal
(Becket, 1989; McLaughlin et al., 2000). It has been suggested that CaCl2 extraction is applicable for testing
bioavailable metals because the extractant does not interfere with the natural pH of the soil, is nondestructive on
most soil minerals, and provides cations for exchange with the metal being tested (Andrewes et al., 1996; Houba
et al., 1996). Theoretically the CaCl2 extractant works by exchanging Ca with metals on the exchange complex
(Houba et al., 1990), thus providing a measure of soil solution plus easily exchangeable metal, i.e. a measure of
immediately bioavailable metal plus the buffering capacity of the soil.
12
As a potential extractant CaCl2 seems to be promising with many studies finding CaCl2 extracted metal to
correlate well with plant responses. Calcium chloride extracted Al was found to correlate well with T. aestivum
root elongation in 20 soils (r2=0.80) (Wright et al., 1989) was the best extractant for T. aestivum in an acid red
podzol (r2=0.92) compared to soil solution (r2=0.78) and BaCl2/NH4Cl extraction (r2= 0.87) (Noble and Randall,
1998). In eight soils with a range of Cd contamination, Latuca sativa (lettuce) growth was found to correlate
well with CaCl2 extracted Cd (r2=0.89-0.91 dependent on extractant concentration) (Andrewes et al., 1996).
Cadmium in Trifolium subterraneum (subterranean clover) plants correlated well (r2=0.91) with CaCl2 extracted
Cd across 3 soils spiked with differing amounts of HCl/ CaCO3 to achieve a range of Cd availabilities (Whitten
and Ritchie, 1991). In four soils, each contaminated with 10 rates of sewage sludge, both CaCl2 extraction and
soil solution gave the best relationships for Hordeum vulgare (barley), Beta vulgaris (beet) and Trifolium repens
(white clover) uptake of Zn, Cu and Ni (Sanders et al., 1987). In a similar study, while CaCl2 was found to be
the most appropriate test when compared with DTPA, EDTA, and soil solution, it only gave consistently high
correlations for Zn (r2>0.67) compared with Ni, Cd and Cu (many r2<0.34) (Sanders et al., 1986).
Differences in methodology are common, with variations existing in the shaking time, soil to solution ratio and
CaCl2 concentration (Esnaola et al., 2000). While differing concentrations of CaCl2 have been used as an
extractant, it is suggested that 0.01M is the most appropriate concentration because of the widespread usage of
this concentration for pH measurements (Andrewes et al., 1996) and the similarity with the ionic strength of
many soil solutions (Houba et al., 2000). However, it has been suggested that 0.01M CaCl2 might not be a
strong enough extractant to differentiate at marginal levels of toxicity (Esnaola et al., 2000).
3.2.5 Soil solution and pseudo-measures of soil solution
The concentration of a metal in the soil solution is a measure of the immediately bioavailable metal (Pearson,
1971). Soil solution is technically defined as the aqueous liquid phase of soil at field conditions (Wolt, 1994).
Thus, true measures of soil solution occur at moisture levels of field capacity or less. Methodologies that
incorporate saturated extracts and water extracts, while included in this discussion, are only pseudo-measures of
soil solution and have been previously called “operationally defined soil solution” (Wolt, 1994). Studies have
regularly found good correlations between soil solution metal concentrations and factors affecting metal
availability. For example, soil solution Cu, Zn and Cd have been found to correlate well with pH, organic
matter, and total metal concentration in the soil (McBride et al., 1997).
Some studies have shown that metal concentration in the soil solution correlates well with plant growth, however
the positive relationship is not conclusive. Vigna unguiculata foliar Mn concentration showed a log linear
relationship with soil solution Mn with an r2 of 0.89 in comparison to DTPA extracted Mn which had an r2 of
0.54 (Vega et al., 1992). As stated in section 3.2.4, both soil solution and CaCl2 extraction gave the best
relationship for uptake of Zn, Cu and Ni in three species grown in four soils contaminated with ten rates of
sewage sludge (Sanders et al., 1987). Soil solution Cu has been shown to have a poor relationship with root,
shoot and total plant Cu in P. vulgaris (Minnich et al., 1987). On three soils regressed independently, soil
solution Al correlated well with Macadamia integrifolia (macadamia) Al uptake but Mn relationships were best
described by a “response area” rather than a single curve (Hue et al., 1987). However the need to regress soils
independently in that study, suggests a lack of robustness in soil solution metal as a measure of bioavailable
metal. In an experiment on ten contaminated soils, Raphanus sativus (radish) tissue Cd correlated well with the
Cd in the soil solution (adjusted R2>0.90) but tissue Zn did not correlate well with Zn in the soil solution
(adjusted R2=0.25-0.53) (Lorenz et al., 1997). Addition of total soil Zn to the regression improved the
relationship (adjusted R2=0.85) suggesting a role for replenishment factors in Zn availability. However, it
should be noted that relationships were limited to linear regressions and the possibility of nonlinear relationships
was not assessed.
At present there is no standardised method of measuring soil solution. Methods from leaching to centrifugation
(Jeffery and Uren, 1983) and differing ratios of water to soil (Minnich et al., 1987; Fotovat et al., 1997) have
been used. Increasing the water to soil ratio when extracting soil solution in non-contaminated soils has been
found to decrease the concentration of Cu in the resulting solution but to increase the Cu extracted per unit
weight of soil (Fotovat et al., 1997). This means that differences between studies and findings could just as
easily be attributable to soil solution extraction technique as to differences between soils or plants. More
importantly, differences in soil solution determinations could be responsible for differing abilities of soil solution
metal to predict plant available metal across studies.
There are two main reasons why total soil solution metal may not be the best indicator of immediately
bioavailable metal. Firstly, metals, and Cu in particular, form complexes with dissolved organic matter in
13
solution thus rendering significant quantities of the metal in the soil solution relatively unavailable (Jeffery and
Uren, 1983). Secondly, there are indications for some metals, as discussed in the next section, that not all metal
species are available to plants, with the most toxic form being the free metal ion.
3.2.6 Speciation and toxic metal forms
The speciation of an element is the chemical form(s) that it takes in solution. Some research suggests that the
free metal ion is the most toxic form of metals. Free metal activity in a wide variety of soils and with a variety
of metals correlates well with factors such as organic matter and pH, suggesting that it is a good measure of plant
available metal (Reddy et al., 1995; Sauve et al., 1997). Research with plants, especially in solution culture, has
supported the free metal ion hypothesis (Sparks, 1984; Bell et al., 1991; Ibekwe et al., 1998). Original
suggestions that not all metal in solution was available to plants began with the findings of DeKock (1957) on
the effect of chelators on uptake of metals. Subsequent work with Al has further supported this theory (Alva et
al., 1986; Menzies et al., 1994).
The “free ion as the bioavailable species” theory has often been assumed in the literature with very little rigorous
research to ground the hypothesis (Parker and Pedler, 1997). Indeed, recent work suggests that the free ion
activity model may be a simplification. Work with chelates and organic acids has shown that metals complexed
with organic compounds are more available than the free-ion activity would suggest (Bell et al., 1991; Laurie et
al., 1991b; Laurie et al., 1991a; Srivastava and Appenroth, 1995; Parker et al., 2001). At a given activity in
solution, plant uptake of Zn and Cd has been found to be greater in the presence of a ligand compared to in the
absence of a ligand, and that this difference was amplified as the metal-ligand binding constant increased
(McLaughlin et al., 1997a). Work with Cd in both soil and solution culture has shown that Cl complexed Cd is
also available to plants (Smolders and McLaughlin, 1996; Smolders et al., 1998). In contrast, research with
Kandelia candel (mangrove) has suggested that NaCl reduces Zn and Cu toxicity (Chiu et al., 1995) but from the
experimental design this could just as easily indicate competition between Na+ and Zn2+/Cu2+ for uptake or a
positive effect of NaCl on growth in this halophytic species. Research with Cu toxicity has shown that plant
responses cannot be explained by the Cu2+ ion alone (Lexmond and van der Vorm, 1981). Parker and Pedler
(1997) conducted computer simulations of various nutrient systems and found that while the free ion activity is
generally a good measure of plant available metal there are certain situations outside the bounds of the model.
For instance, use of the free ion activity as the measure of plant available metal assumes that the solution is not
replenished from solid phases, while this holds in most solution cultures it does not in soil studies. They also
found that larger deviations from the model were likely to occur at high root to solution ratios, as in soil.
Research between metals and the free metal ion in soil solution is inconclusive on whether the free metal ion is
the best predictor of metal availability. In R. sativus, L. sativa and Lolium perenne (perennial rye grass) free ion
activities were better measures of plant Cu uptake than total Cu or CaCl2 extracted Cu in eight contaminated
urban soils (Sauve et al., 1996). Research with T. subterraneum and Panicum virgatum (switch grass) found
plant uptake of Mn to be equally well correlated with Mn2+ activity and the soil solution Mn concentration across
11 soils (Wright et al., 1988). The Cu2+ activity in soil solution has been found to not predict P. vulgaris Cu
uptake (Minnich et al., 1987). Concentration of Cd in Solanum tuberosum (potato) tubers from 50 soils was not
related to the Cd2+ activity in the soil solution but rather to the degree of Cd-Cl complexation in the solution,
suggesting a role for more than the free ion in Cd toxicity (McLaughlin et al., 1997b). While the debate
continues, evidence suggests that the immediately bioavailable fraction of a metal is somewhere between the
total amount in the soil solution and the free ion activity. Until the answer is found for each metal either the free
ion activity or total soluble metal in solution appear to be the best estimates.
3.2.7 Plant factors
While soil factors have a large impact on the bioavailablity of metals to plants, different species or varieties
grown on the same soil can have different metal uptakes (Miles and Parker, 1979). Therefore, there are species
specific factors affecting plant uptake. Jarvis and Whitehead (1981) suggested that a true measure of plant
available metals will not be attained unless the extent of soil exploitation by the roots is accounted for.
However, it should be noted that while two plant species may take up a different amounts of metal within a given
time frame it does not necessarily mean they are extracting from different soil pools of metal.
3.2.8 Conclusions on the bioavailability of metals in soils
Available data suggest that the immediately bioavailable fraction of a metal is somewhere between the total
amount in the soil solution and the free ion activity. Potential exists for tests such as CaCl2 extraction and ion
exchange resin techniques which measure the immediately bioavailable metal plus the buffering capacity of the
soil solution. However, tests which measure the immediately bioavailable metal plus longer term replenishment
14
(e.g. DTPA and EDTA) do not correlate well with plant responses across a range of soil types. Within the tests
holding promise there is a strong need for standardisation of methodology, good quality testing, and comparisons
of all promising soil tests across a wide range of contaminated soil types and against plant responses for multiple
species. The explanation may include factoring in parameters such as the relative toxicity of different metal
species, the ameliorative qualities of other ions, e.g Ca2+ (Vega et al., 1992; Kinraide, 1998), the ionic strength
of the nutrient/soil solution (Kelly et al., 1990), the effect of organic matter (Sims, 1986), and possibly plant
factors (Jarvis and Whitehead, 1981).
3.3 Relationship between soil and solution culture studies
3.3.1 Solution culture as a model of the soil (solution)
A “good model” is a simplification of a natural system that is able to simulate the major factors of that system.
Hence, nutrient solutions can be considered models of the soil system or more specifically of the soil solution.
Whether or not a nutrient solution is a good model of the soil (solution) depends on experimental design.
The main advantage that solution culture has over soil culture is that the composition of the growing medium can
be defined, manipulated, and measured with a high level of precision (Rorison and Robinson, 1986). This is
important when undertaking research into metal toxicity. The number of factors affecting metal availability,
issues over what metal fraction is available, and the large amount of variability in soil, make toxicity studies very
difficult if quantitative and nonconfounded results are to be obtained
3.3.2 Comparing toxicity in soil with that in solution culture
While dilute solution culture studies offer a relatively good method of approximating the soil solution (Asher
and Edwards, 1978; Sparks, 1984; Parker and Norvell, 1999) it is still a simplification of the soil system. The
main difference between solution culture and soil, is that in solution culture plant roots are bathed in a
continuously aerated and mixed aqueous medium, while in soil where plant roots must continually explore the
soil to gain nutrients and water. Therefore, it is likely that in soil any toxic metal induced reduction in root
growth would have a larger effect on overall plant biomass than in solution culture (Patterson and Olson, 1983).
It also seems likely that this effect would be somewhat larger than the decrease in biomass as the larger roots
make up most of the biomass, meaning that lateral and fine roots could be affected before an appreciable
decrease in root biomass became apparent. This is in keeping with the results of Brown and Wilkins (1985a)
who found root elongation was more sensitive than root biomass to Zn in Betula sp. (birch). In addition, many
metal toxic soils have lower-than-optimal nutrition, low pH, and combinations of toxic metals; thus intensifying
the effect of individual toxic metals further (Godbold and Huttermann, 1985).
Research has shown that fertilised soils result in lower concentrations of toxic metals in the plant tissues (Merry
et al., 1986a). This may be because of competition for uptake at the root surface or because of increased growth
resulting in dilution of the metal in the plant’s tissues. The effects of ionic strength have also been seen in
solution culture experiments (Vlamis and Williams, 1962; Brown and Wilkins, 1985a). As such, nutrient
solutions with high ionic strengths could only be comparable to highly fertilised and/or very fertile soils. This
highlights the importance of using dilute nutrient solution cultures.
In soil, plants may experience metal concentrations that vary considerably in space and time. In comparison,
nutrient solution metal concentrations are kept relatively constant through out time and the only spatial variation
in well mixed nutrient solutions might be at the root-rhizosphere-bulk solution interfaces. Little research has
been undertaken to compare the effects of variability of metal concentration on growth. Tack et al. (1998) have
recently found that it is the average Cd concentration in nutrient solutions which is important for plant uptake
and that varying the way in which the Cd concentration fluctuated in the short term had no significant effect.
They also found that nutrient solution concentrations towards the end of growth were the most important. Both
of these findings suggest that soil solution metal readings taking towards the end of the growth period should
correlate well with analogous average metal concentration in nutrient solutions.
Interaction effects of metals on plant growth in solution culture can be different to those found in soil. Luo and
Rimmer (1995) found that Cu-Zn interactions in a soil were synergistic compared with an antagonistic effect in
solution culture. Luo and Rimmer (1995) attribute this difference to the effects on availability of metal
adsorption onto soil particles. Taylor et al. (1998) suggest that these differences are likely to be because most
nutrient solution studies are conducted at high ionic strength compared to soil solutions.
15
4. Plants and metals
4.1 Plant uptake and transport of metals
Plants have developed a range of mechanisms to obtain metals from the soil solution and transport these metals
within the plant. Much of the research, and hence understanding, of these mechanisms has been at sufficiency
and deficiency levels of metals. However, from an understanding of the mechanisms operating at deficiency and
sufficiency levels of metals, supplemented with what is understood at excess metal supply, an understanding can
be gained of the processes affecting metal uptake and transport by plants.
4.1.1 Uptake mechanisms
Uptake of metals into plant roots is a complex process involving transfer of metals from the soil solution to the
root surface and inside the root cells. Understanding of uptake processes is hampered by the complex nature of
the rhizosphere which is in continual dynamic change interacted upon by plant roots, the soil solution composing
it and microorganisms living within the rhizosphere (Laurie and Manthey, 1994).
4.1.1.1 Supply of bioavailable metals for plant uptake.
The bioavailable fraction of metals in a soil is generally thought to be the free metal ion in the soil solution.
However, as discussed in section 3.2.6, this appears to be an over simplification in some circumstances. In
deficiency and sufficiency situations the free metal ion activity in the soil solution is low and plants have
developed strategies to maximise the potential uptake of metals (Welch, 1995). Plants are able to influence the
solubility and speciation of metals in the rhizosphere by exuding chelators (Fan et al., 1997) and manipulating
rhizosphere pH.
Most of our understanding of plant metal uptake has come from the study of Fe (Kochian, 1993). For Fe uptake,
two different strategies have been identified. In Strategy I plants, i.e. dicots and nongraminaceous monocots,
Fe(III)-chelates or -complexes present in the rhizosphere are reduced by plant produced reductants in the
rhizosphere for uptake with other sources of free Fe2+ across the plasma membrane (Chaney et al., 1972; Brown
and Ambler, 1973; Olsen and Brown, 1980; Welch et al., 1993). In addition, plant produced organic acids are
excreted which can complex with Fe (Grusak et al., 1999). It does not appear as though Strategy I plants are
able to directly absorb Fe-chelates or Fe-complexes (Chaney et al., 1972). In addition to the above mechanisms,
Strategy II plants, i.e. graminaceous monocots, excrete chelates such as mugeneic and avenic acids (Kochian,
1993; Fan et al., 1997) which are known as phytosiderophores or phytometallophores depending on their
association with Fe alone or all metals respectively, into the rhizosphere (Fan et al., 1997). Iron, and other
metals, chelate with the phytometallophores, providing a ready supply of metals for reduction and transport
across the plasma membrane. Some research has also shown that Strategy II plants can directly absorb the Fephytometallophore complex (Grusak et al., 1999).
Phytometallophores complex other metals as well as Fe (Treeby et al., 1989), and so, it is generally assumed that
uptake of other metals will be similar to Fe although conclusive evidence is lacking (Grusak et al., 1999). Under
Fe deficient conditions H. vulgare plants absorbed extra Cu, Mn and Zn suggesting a multiple role for
phytometallophores in metal uptake (Fan et al., 1997). However, Zn deficiency in the same species only resulted
in a small increase in phytometallophore release when compared to Fe deficiency, and deficiencies of Cu and
Mn did not result in any increased phytometallophore release (Gries et al., 1995). Zea mays (maize) roots
released phytometallophores under conditions of Zn deficiency forming Zn-phytometallophore complexes in the
rhizosphere (Von Wiren et al., 1996). This research also demonstrated that doubly radiolabelled 65Zn(14C)phytometallophore complexes were directly absorbed (Von Wiren et al., 1996). In another study with Z.
mays roots, doubly radiolabelled 64Cu-(14C)tetrethylenepentamine was directly absorbed (Smeulders et al.,
1983). Hence, the information is inconclusive on the importance of phytometallophores in Cu, Mn and Zn
nutrition of Strategy II species. The studies do however suggest that Strategy II species are capable of the direct
absorption of a range of metal-chelates, however higher uptakes of the free metal ion compared to the metalphytometallophore suggest a preference for the free ion (Von Wiren et al., 1996).
There is less agreement for the uptake of other metals besides Fe in Strategy I species, not least because only
limited amounts of research have been undertaken. Iron deficiency increased uptake of Mn in Glycine max
(soybean) and the authors suggested this occurred because of the Fe deficiency induced increase in reductants
mobilising extra Mn (Baxter and Osman, 1988). Copper depleted P. sativum increased reduction of both Cu and
Fe compared to controls in the same area of the root where Fe deficient plants increased reduction of Fe, thus
suggesting a similar mechanism for Cu and Fe uptake (Welch et al., 1993). However, further research with
16
Pisum sativum (pea) has suggested that free Cu2+ rather than Cu-chelate was the substrate for reduction by the
plasma membrane for transport into root cells (Holden et al., 1995). It was only under conditions of excess Cu
that the Cu-chelate complex seemed to be the substrate for plasma membrane Cu reduction (Holden et al., 1995).
This finding is counterintuitive, for, one would suspect that at excess metal conditions, where higher
concentrations of free metal ions were present, the plant would be unlikely to commence reduction of metalchelates. Holden et al. (1995) did not suggest a reason why plants would access an extra pool of soil Cu under
conditions of excess Cu. Possibly the results indicate a breakdown in normal functioning under toxicity
conditions.
In addition to reduction of chelated metals to increase the free metal supply there is also limited evidence for the
direct reduction of precipitated metal, suggesting a further supply of metals within the rhizosphere. For example,
roots of many species, including Strategy I and II species, have been shown to directly reduce Mn-oxides to
Mn2+ (Uren, 1981).
Few studies have been concerned with the role of phytometallophores under conditions of excess metal.
However, as phytometallophore release has been associated with deficiency rather than sufficiency conditions it
is clear that plants can regulate phytometallophore release and are unlikely to release increased concentrations of
phytometallophores at excess metal supply. One study, which dealt with phytometallophore release under Fe
deficiency induced by excess Mn supply, demonstrated that the induced Fe deficiency caused an increase
production of phytometallophores (Alam et al., 2000). This increase was less than in plants experiencing Fe
deficiency without excess Mn, suggesting that under metal toxic conditions release of phytometallophores is
reduced by the plant. Hence, under conditions of excess metal, plant produced chelates are likely to play a
smaller role in the supply of metals than in deficiency situations. Instead, under conditions of excess metal, the
direct supply of free-metal ions, which would be in greater supply than in deficient situations, and metal-organic
complexes produced by decaying biomatter and microbes in the rhizosphere would have a proprotionately
greater influence on metal uptake.
4.1.1.2 Manipulation of rhizosphere pH
The pH of the rhizosphere may vary by up to 2.5 pH units from that of the bulk soil solution depending on plant
species, plant age, nutrient supply and the buffer capacity of the soil (Hedley et al., 1982; Marschner et al.,
1982; Pilet et al., 1983; Romheld et al., 1984). This is mainly as a result of an imbalance in cation/anion
uptake and, hence, excretion of H+/OH- (or HCO3-), excretion of organic acids, production of CO2 and microbial
activity in the rhizosphere (Hedley et al., 1982; Clarkson, 1985; Marshner, 1993). Hence, the solubility,
speciation and corresponding availability of metals in the rhizosphere may be different to that in the bulk soil
solution. Plants, especially strategy I species, are known to acidify the rhizosphere when the supply of nutrients
are deficient thus increasing the availability of metals (Olsen and Brown, 1980; Marschner et al., 1982;
Romheld et al., 1984; Marschner et al., 1986; Brown and Jolley, 1988). In addition, as discussed further in
section 4.6.1.1, some authors have suggested that plants growing in situations of excess metal supply can
increase rhizosphere pH to minimise metal availability. However, there is little conclusive evidence to support
this hypothesis.
4.1.1.3 Role of mycorrhiza
Mycorrhiza are mutualistic associations between certain soil fungi and the roots of most plant species (Reddell
and Milnes, 1992; Brundrett et al., 1996). The mycorrhizal fungi benefit from the association by obtaining
photosynthates and, in exchange, mycorrhizal fungi increase the plant uptake of P (Jasper et al., 1988; Burgess
et al., 1993) and trace metals (Clarkson, 1985; Pahlsson, 1989). Mycorrhizal fungi achieve this increase in plant
nutrition by increasing the surface area of soil explored compared with non-mycorrhizal roots (Clarkson, 1985)
and increasing the solubility of metals e.g. by producing metal-chelators (Szaniszlo et al., 1981).
It has been suggested that as well as assisting nutrient uptake at low metal concentrations, mycorrhizal fungi are
able to reduce metal uptake, or at least increase plant metal tolerance by affecting metal translocation, under
conditions of metal contamination. Some studies have found this to be the case, for example, mycorrhizal
Trifolium pratense (red clover) plants grown in acid soils had less Mn in the roots and the shoots than nonmycorrhizal plants (Arines et al., 1989). Other studies have found little change in uptake but a decrease in metal
translocation to the shoots resulting in, for example, increased tolerance of Betula sp. to excess Zn (Brown and
Wilkins, 1985b) and Vaccinium macrocarpon (cranberry) to excess Mn (Hashem, 1995). However, a critical
review of the literature finds many conflicting and negative results (Killham and Firestone, 1983; Jones et al.,
1986; Pahlsson, 1989). It appears that the ability of a mycorrhizal fungi to increase plant metal tolerance is
affected by other growth conditions, the fungal species and the metal type (Kahle, 1993; Weissenhorn et al.,
17
1995). Of high importance is the tolerance of the mycorrhizal fungi to excess metal as the plant could be more
tolerant than the fungi (e.g.Arines and Vilarno, 1991).
4.1.1.4 Uptake of metals across the plasma membrane
Metal uptake by plants is regulated by the electrochemical potential gradient for each metal ion that exists across
the plasma membrane of root cells (Welch, 1995). Most plants have a plasma membrane potential between –120
and –180 mV, hence a large electrical gradient exists that powers metal uptake (Kochian, 1991; Welch, 1995).
As well, the metal ions in the cytoplasm are maintained at low activity to prevent harmful redox reactions which
can result from the presence of free ionic forms of these reactive metals (Laurie and Manthey, 1994; Welch,
1995). These two factors combine to create a large passive gradient for metal uptake. Hence, as opposed to
macronutrients, there is little need for the plant to utilise thermodynamically active processes for the uptake of
metal ions (Kochian, 1993; Welch, 1995; Von Wiren et al., 1996).
The exact nature of the membrane transporter, which controls influx across the plasma membrane into the
cytoplasm, is not yet known. It has been suggested that the transporter may be a metal specific or nonspecific
channel protein (Kochian, 1993; Welch, 1995; Grusak et al., 1999) although conclusive evidence does not
exist. It appears that the capability of the root membrane transport mechanism is far in excess of the plant metal
requirements (Welch, 1995; Grusak et al., 1999) indicating the mechanism by which toxic concentrations of
metals may enter the plant roots.
4.1.2 Transport within the plant
Once within a plant the two major transport mechanisms for metals are via the xylem and phloem. The effects of
metals on the rate of movement and composition of the xylem and phloem sap may impact on plant response to
metal toxicity. Translocation effects include the relative proportions of metals in roots vs. shoots, potential sites
of toxic action of metals and the translocation of other nutrients within the plant
4.1.2.1 Xylem transport
Transport of metal ions within the xylem is essentially driven by mass upward flow of water created by the
transpiration stream (Kochian, 1991; Welch, 1995). Water transpiration rate has a large effect on macronutrient
translocation rate, however at low supply, processes such as xylem loading and unloading and transfer between
xylem and phloem have been shown to be more important for the rate of nutrient supply (Welch, 1995). There is
little to suggest that the case would be different for metals, and hence, under conditions of excess metal it is
likely that the rate of transpiration would dominate metal movement in the xylem sap.
In addition, the composition, pH and redox potential of the xylem sap would effect the types and amounts and
therefore movements of metal species in the xylem sap (Welch, 1995; Liao et al., 2000b). Both xylem pH and
redox potential are important for regulating the solubility and speciation of metal within the xylem (Welch,
1995), hence effecting the concentration that can be transported throughout the plant. Plants appear to be able to
maintain xylem pH (5.4-6.5) (Hocking, 1980; Marschner, 1995) and redox potentials (Welch, 1995) at fairly
constant levels at deficiency and sufficiency levels. No comprehensive research exists on the effects of excess
metal on xylem pH, redox potential or ionic strength, however, the little which does exist suggests plants are able
to maintain these characteristics (White et al., 1981c; Clarkson et al., 1984).
Copper in xylem sap has been shown to be almost 100% bound to amino acids and this high percentage of
complexation is retained under conditions of excess Cu supply (Graham, 1979; White et al., 1981b; Pich and
Scholz, 1996; Liao et al., 2000b; Liao et al., 2000a). In addition, in both Lycopersicon esculentum (tomato)
and Cichorium intybus (chicory) increasing the supply of Cu increased the production of amino acids,
particularly nicotianamine and histadine (Liao et al., 2000b). This suggests that even under toxic conditions
plants have mechanisms to regulate complexation of Cu within the xylem sap and hence minimise potential
damage caused from high concentrations of free Cu ions (Welch, 1995).
At both sufficiency and toxic concentrations, computer simulations run on xylem sap composition suggest that
Zn is predominantly complexed to citric and malic acid in L. esculentum and G. max (White et al., 1981a). At
excess Zn, small amounts of soluble Zn-phosphate were found in the sap (White et al., 1981a). As well, for G.
max increasing the Zn supply increased the concentration of free Zn2+, the opposite being true for L. esculentum
(White et al., 1981a). Hence, while plants appear able to adjust to excess Zn in the xylem, there are indications
in some species that an increasing concentration of potentially harmful Zn2+ could occur. In contrast, excess Zn
does not appear to affect the speciation of Cu in the xylem (Liao et al., 2000b).
18
Compared with Cu and Zn, a greater proportion of the Mn in the xylem fluid exists in the uncomplexed, free-ion
form. Thus, in computer speciation studies, 37% of the Mn in G. max xylem sap and 72% in L. esculentum
xylem sap was found as Mn2+ (White et al., 1981a). Manganese was also found complexed to organic acids
rather than amino acids in both species (White et al., 1981a). In another study, Mn in the xylem exudate of
Helianthus annuus (sunflower) was mainly present as Mn2+ at Mn supply ranging from deficient to toxic
(Graham, 1979). Hence, the opportunity exists for high concentrations of harmful Mn2+ in the xylem fluid at
excess Mn supply.
Copper, Mn and Zn are predominately transported throughout the plant within the xylem rather than the phloem
(Pearson et al., 1996). Exudation rates of Zn have been found to be decreased under excess Zn in L. esculentum
but not G. max (White et al., 1981c). Xylem exudation rate was reduced in L. esculentum when excess Mn was
supplied (Le Bot et al., 1990). This suggests that in some species metal toxicity may have an effect on the rate
of xylem transport, possibly by reducing transpiration (Rousos et al., 1989; Brune et al., 1994). This could have
effects on the concentrations of other nutrients reaching the shoots.
Under conditions of elevated metal supply, generally the majority of metals are restricted to the plant roots
(Brown and Wilkins, 1985a; Kelly et al., 1990; Leita et al., 1993; Harrington et al., 1996). It is likely that this
occurs because of some unknown mechanism which prevents xylem loading of excess metals. It has been
suggested that this mechanism may incorporate a high number of metal specific binding sites in the roots (Liao
et al., 2000a). This could explain why plants grown under conditions of excess Mn regularly have a greater
proportion of shoot Mn (Pettersson, 1976; Chino and Baba, 1981), as Mn has a low binding capacity (McBride,
1982; Norvell, 1991) and hence is less likely to be restricted from entering the xylem, compared to Cu and Zn.
4.1.2.2 Phloem transport
Transport of metals within the phloem is thought to occur via the positive hydrostatic pressure gradient
developed from the loading of sucrose into the phloem from mature actively photosynthesizing leaves and
unloading of sucrose into the sink tissues such as rapidly growing tissues, apical root zones and reproductive
organs (MacRobbie, 1971; Hocking, 1980; Welch, 1995). As in the xylem, the pH, redox potential, ionic
strength and organic constituents of the phloem sap will determine the loading, transport and unloading of metals
in the phloem (Welch, 1995). However, unlike xylem cells, phloem cells are alive and metabolically active.
Hence, metabolic reactions within the phloem have the potential to make the phloem sap more responsive to
changes in the internal plant environment than the xylem sap (Welch, 1995).
Typically phloem sap has of a pH of 8 or greater (Hocking, 1980; Kochian, 1991), is more reducing and has a
higher solute concentration (Hocking, 1980) and ionic strength (Welch, 1995) than xylem sap. Hence, the
activity and speciation within the phloem sap is likely to be considerably different to the xylem sap. Sugars have
been found to compose from 14 to >24% of the phloem sap (Hocking, 1980). No comprehensive research has
studied the effect of excess metals on phloem composition and so it is not known whether plants are able to
maintain a stable pH, redox state and ionic strength under excess metal supply.
Information on the speciation of Cu, Mn and Zn within the phloem is scarce (Stephan and Scholz, 1993). At
sufficiency levels, Zn in the phloem sap of Ricinus communis (castor bean) was almost all bound to organic
molecules (molecular weight between 1000-1500) with small amounts present as Zn2+ (Van Goor and Wiersma,
1976). In comparison, Mn was 60-70% present in ionic forms with the remainder bound to organic molecules
(molecular weight between 1000 and 5000) (Van Goor and Wiersma, 1976). Information does not exist on the
speciation of Cu within the phloem, however because of the propensity of Cu to form complexes with organic
molecules (Norvell, 1991) it is likely that little would exist as the free-ion. Nicotaniamine has been suggested as
a transport molecule for Cu, Mn and Zn in the phloem (Stephan and Scholz, 1993) but conclusive evidence is
lacking.
Copper is classified as having variable phloem mobility dependent on plant species and Cu status of the whole
plant as well as the source and sink organs (Stephan and Scholz, 1993; Welch, 1995). Excess Cu has been
shown to reduce phloem transport of Zn but not Mn into T. aestivum grains (Pearson et al., 1996). The authors
suggested that this was because Cu and Zn competed for the same phloem loading sites.
Manganese mobility within the phloem is generally considered to variable and is dependent on the Mn status of
the plant species as well as the source and sink organs (Welch, 1995; Pearson et al., 1996). At sufficient Mn
supply in T. aestivum, only a small proportion of Mn was transported in the phloem (Herren and Feller, 1997a).
Under conditions of excess Mn supply, little effect was noted on phloem mass flow as evidenced by a lack of
change in sucrose transport within T. aestivum (Pearson et al., 1996). In contrast, excess Mn supply reduced
19
phloem transport of Mn and the authors suggested loading from the xylem into the phloem may have been the
rate limiting process (Pearson et al., 1996).
Zinc is considered to have variable phloem mobility dependent on the Zn status of the plant species as well as
individual plant tissues and organs (Welch, 1995; Pearson et al., 1996; Herren and Feller, 1997a; Herren and
Feller, 1997b). In a split-root experiment at inadequate Zn supply, Zn supplied to one section of roots was not
transported to the other root section suggesting a lack of Zn transport in the phloem (Welch et al., 1999). In
contrast, at sufficient Zn supply in one root section an average of 29% of the Zn taken up was transported to the
other root section suggesting an increasing role for phloem transport as Zn supply increases (Welch et al., 1999).
In contrast, in T. aestivum excess Zn supply has little effect on phloem mass flow, as evidenced by rate of
sucrose transport, but resulted in a reduction in phloem transport of Zn probably as a result of reduced phloem
loading of Zn (Pearson et al., 1996; Herren and Feller, 1997a). This could indicate a stress response or suggest
a tolerance mechanism to prevent excess Zn redistribution to young developing tissues (Herren and Feller,
1997a). In contrast, under conditions of excess Zn supply a reduced phloem loading of sucrose was noted in P.
vulgaris (Rauser and Samarakoon, 1980). In addition, excess Zn supply has been shown to have a negative effect
on phloem loading and transport of other metals (Herren and Feller, 1996; Welch et al., 1999) thus having the
potential to induce local deficiencies and/or toxicities. Rauser and Samarakoon (1980) thought that a Zninduced limitation in the supply of ATP might be responsible for reduced phloem loading, and thus transport,
under conditions of excess Zn and could be responsible for much of the toxic effects of excess Zn on plant
biomass production.
4.2 Relationship between metal in the growth substrate and growth
Plant responses to metals are dose dependent (Berry and Wallace, 1981). For essential metals these responses
cover the phases from deficiency through sufficiency/tolerance to toxicity. In non-essential metals only the
tolerance and toxicity phases occur.
The idea of a critical or threshold toxicity is often used to establish the point at which metals in the growth
substrate cause significant growth decrease. These are often defined as the substrate metal concentration
corresponding to a yield decrease of 10%. However, the point at which the level tolerance plateau changes to the
steep decrease of the toxicity phase (i.e. the point of inflexion) is also commonly used (Davies, 1993).
While the idea of critical concentration for toxicity is valuable, it relegates toxicity effects to one point along the
response curve. Berry and Wallace (1981) use the idea of unit toxicity as well as threshold toxicity to compare
plant responses. Unit toxicity is defined as the yield decrease per unit increase in metal in the substrate after the
threshold toxicity is passed. From combining these two ideas not only can the point at which toxicity
commences be established but also the rate at which toxicity progresses. A more easily determined measure is
the metal concentration corresponding to a 50% decrease in yield (e.g. Zhang et al., 1998). From a comparison
of the metal concentrations corresponding to a 10% and a 50% decrease in yield a qualitative estimate of the rate
of toxicity can be determined.
4.3 Relationship between tissue concentration and growth
The idea of a critical or threshold tissue concentration relating to the commencement of a significant growth
decrease as a result of toxicity is regularly used for diagnosis of toxicity and to compare tolerances between
species and metals. A common method of defining the critical tissue concentration is the tissue concentration
resulting in a 10% reduction in growth or yield (Edwards and Asher, 1982; Aery and Jagetiya, 1997; Zhang et
al., 1998). Dry matter yield decrease has generally been accepted as the standard measure for comparisons of
toxicity. However, occasionally other measures such as fresh weight, commencement of symptoms (Elamin and
Wilcox, 1986b), and metabolic responses have been used (Gherardi et al., 1999). Traditionally, the youngest
fully expanded leaf (YFEL) has been the plant tissue of choice for comparisons in nutrient studies. Within a
species, YFELs should be of a similar age and therefore will have been taking up nutrients for a similar time
irrespective of whole plant age and so, theoretically, should allow for direct comparisons between relatively
young research material and older field or wild grown plants. Hence, YFEL tissue concentrations are seen as
preferable to whole shoots or whole root concentrations where concentration varies with plant age. However,
while much research in general plant nutrition has utilised the YFEL as an index leaf (Asher et al., 1984; Dell
and Daping, 1995; Gherardi et al., 1999) usage is not wide-spread in toxicity work. From Table 4.1, Table 4.2
and Table 4.3 it can be seen that the critical tissue concentrations for toxicity vary considerably across metal,
species, and the plant tissue being measured.
20
Table 4.1 Critical tissue concentrations for Cu toxicity in a variety of species
Species
Plant part
Twelve spp. of tropical
grass
A. hypogaea
Carthamus tinctorius
(safflower)
G. max
O. sativa
Shoots
Critical toxicity
(mg.kg-1)
181
Soil pot experiment
Shoots
Leaves
2302
101
Soil pot experiment
Sand culture
Method of cultivation
Reference
(Plenderleith and Bell,
1990)
(Borkert et al., 1998)
(Pandey and Sharma,
1999)
(Borkert et al., 1998)
(Lidon and Henriques,
1992)
(Borkert et al., 1998)
(Wheeler and Power,
1995)
(Kalyanaraman
and
Sivagurunathan, 1993)
Soil pot experiment
Shoots
1402
Whole
353
Solution culture
plant
O. sativa
Shoots
<202
Soil pot experiment
T. aestivum
Shoots
754
Solution culture
300
Roots
Vigna mungo
Leaves
671
Soil pot experiment
(blackgram)
Stems
50
Roots
41
1
Tissue concentration at 90% of maximum yield
2
Intersection of regression lines: “shoot growth independent of leaf metal concentration” and “growth inhibition
at higher metal levels”
3
Nutrient calibration curve of for total tissue concentration and root growth plotted on a log-log graph
4
Estimated as a narrow range of plant concentrations where relative yield decreases rapidly
Table 4.2 Critical tissue concentrations for Mn toxicity in a variety of species
Species
Nine spp. of
pasture legumes
Thirteen crop and
pasture spp.
Twelve spp of
pasture legumes
Citrullus lanatus
(watermelon)
Cucumis melo
(muskmelon)
G. max
L. esculentum
Plant part
Leaves
Critical toxicity
(mg.kg-1)
190 (Medicago sativa,
lucerne) to >2000 (T.
subterraneum)1
182 (M. sativa) to >1200
(T. subterraneum)
200 (Z. mays) to 5300
(H. annuus)1
380 (M. sativa) to 1600
(Centrosema pubescens)2
13243
Leaves
9003
Shoots
Young
tissue
Shoots
Shoots
Method of cultivation
Reference
Sand culture
(De Marco et al., 1995)
Solution culture
(Edwards and Asher,
1982)
(Andrew and Hegarty,
1969)
(Elamin and Wilcox,
1986b)
(Elamin and Wilcox,
1986a)
(Ohki, 1976)
(Amberger and Yousry,
1988)
(Kalyanaraman
and
Sivagurunathan, 1993)
Solution culture
Sand culture
Sand culture
Solution culture
Blade 3
1601
Mature
2504
Solution culture
leaves
Soil pot experiment
V. mungo
Leaves
1441
Stems
120
106
Roots
Z. mays
Mature
35004
Solution culture
leaves
1
Tissue concentration at 90% of maximum yield
2
Tissue concentration at 95% of maximum yield
3
Tissue concentration of treatment at which symptoms were first seen
4
Method not stated
(Amberger and Yousry,
1988)
Table 4.3 Critical tissue concentrations for Zn toxicity in a variety of species
Species
Acacia auriculaeformis
Plant part
Shoots
Critical toxicity
(mg.kg-1)
1431
21
Method of
cultivation
Solution culture
Reference
(Zhang et al., 1998)
(earleaf acacia)
A. hypogaea
H. vulgare
P. vulgaris
P. vulgaris
Leaves
Shoots
Roots
Developing
leaves
Mature leaves
Stems
Roots
Primary leaves
2384
5222
500
1342
Solution culture
95
242
486
2265
Solution culture
Soil
Soil pot trial
(Cox, 1990)
(Aery and Jagetiya,
1997)
(Ruano et al., 1987)
(van Assche et al.,
1988)
Saccharum officinarum Shoots
442
Sand culture
(Chatterjee et al.,
(sugarcane)
1998)
T. aestivum
Shoots
20003
Solution culture
(Wheeler and Power,
Roots
5000
1995)
Typha latifolia
Shoots
7821
Solution culture
(Ye et al., 1998)
1
Tissue concentration corresponding to solution Zn concentration at 90% of maximum yield
2
Tissue concentration at 90% of maximum yield
3
Estimated as a narrow range of plant concentrations where relative yield decreases rapidly
4
From a previously determined critical soil Zn level (Mehlich-1 extraction)
5
Intersection of regression lines: “shoot growth ... independent of leaf metal concentration” and “growth
inhibition at higher metal levels”
4.4 Symptoms and visual evidence of toxicity
The most widespread visual evidence of metal toxicity is a reduction in plant growth as metal toxicity increases.
However, as different metals have different sites of action within the plant, the overall visual toxic response
differs between metals.
4.4.1 Copper
4.4.1.1 Toxicity symptoms
Interveinal foliar chlorosis is a common initial symptom of Cu toxicity (Taylor and Foy, 1985; Zhu and Alva,
1993). Chlorosis was also displayed in Banksia ericifolia (heath banksia), Casuarina distyla (she-oak) and
Eucalyptus eximia (yellow bloodwood) when grown at elevated Cu (Mitchell et al., 1988). The chlorosis often
takes the form of cream or white spots or lesions (Lee et al., 1996a; O'Sullivan et al., 1997). With increasing
exposure, leaf tips and margins can become necrotic (Taylor and Foy, 1985; Yau et al., 1991). In acute Cu
toxicity, leaves may become wilted before eventually becoming necrotic (Yau et al., 1991). Copper toxicity can
be associated with a purpling of foliage (Choi et al., 1996) but this is not apparent in all species (O'Sullivan et
al., 1997).
Copper toxicity has a significant effect on root growth and form, often before any effect on above-ground growth
(Minnich et al., 1987). Patterson (1983) found that the germination of six tree species was less sensitive to Cu
than subsequent root elongation. In dicot seedlings, toxic amounts of Cu result in radicles which are short, blunt
tipped, of dark brown/black colouration (necrotic) and have a disposition to fungal attack (Patterson and Olson,
1983). Citrus paradisi x Poncirus trifoliata (swingle citrumelo) seedlings exposed to excess Cu produce few
new roots and have a thickened tap root (Zhu and Alva, 1993). Thickening of root apicies was also apparent in
Pinus seedlings (Arduini et al., 1995). In Betula papyrifera (paper birch) and Lonicera tatarica (honeysuckle)
seedlings high Cu concentrations have been shown to inhibit the production of root hairs (Patterson and Olson,
1983).
4.4.1.2 Symptoms of induced deficiency
Iron uptake can be decreased by excess Cu (Lexmond and van der Vorm, 1981; Yau et al., 1991; Ouzounidou,
1995). This suggests that the chlorotic symptoms on young leaves of plant experiencing Cu toxicity could be an
induced Fe-deficiency.
There is some contention over whether or not excess Cu decreases or increases Zn content in plants (Luo and
Rimmer, 1995). If excess Cu does have an antagonistic effect on Zn uptake then some of the symptoms of Cu
22
toxicity could be induced Zn deficiency. Zinc deficiency commonly presents on young leaves as chlorosis and
reddening (Davies, 1993; Ren et al., 1993; Lee et al., 1996a).
4.4.2 Manganese
4.4.2.1 Toxicity symptoms
Necrotic brown spotting on leaves, petioles and stems is a common symptom of Mn toxicity (Horst and
Marschner, 1978; Wu, 1994). This spotting starts on the lower leaves and progresses with time towards the
upper leaves (Elamin and Wilcox, 1986a; Elamin and Wilcox, 1986b; Horiguchi, 1988). With time the
speckles can increase in both number and size resulting in necrotic lesions, leaf browning and death (Elamin and
Wilcox, 1986a; Elamin and Wilcox, 1986b). General leaf bronzing and shortening of internodes has been
documented in Cucumis sativus (cucumber) (Crawford et al., 1989). Another common symptom is known as
“crinkle-leaf” it occurs in the youngest leaf, stem and petiole tissue and is also associated with chlorosis and
browning of these tissues (Horst and Marschner, 1978; Wu, 1994; Bachman and Miller, 1995). Roots
exhibiting Mn toxicity are commonly brown in colour (Le Bot et al., 1990; Foy et al., 1995) and sometimes
crack (Foy et al., 1995).
Manganese toxicity in some species starts with chlorosis of older leaves moving towards the younger leaves with
time (Gupta, 1972; Elamin and Wilcox, 1986a; Bachman and Miller, 1995). This symptom starts at the leaf
margins progressing to the interveinal areas and if, the toxicity is acute, progressing to marginal and interveinal
necrosis of leaves (Bachman and Miller, 1995). In the only research on Mn toxicity of Australian native trees,
Eucalyptus gummifera (red bloodwood) displayed small, chlorotic leaves that were often distorted in shape, and
death of terminal buds (Winterhalder, 1963).
4.4.2.2 Symptoms of induced deficiency
The chlorosis displayed by many species under Mn toxicity is similar to Fe deficiency and many studies suggest
that the symptoms are induced Fe deficiency. Nutrient solution containing high levels of Mn have been found to
increase the amount of Fe deposited on the roots surface and a decrease in the amount of Fe translocated to the
shoots (Sideris, 1950). Excess Mn results in a decrease in Fe uptake and transport in G. max (Lingle et al.,
1963). It has been shown that Fe deficiency symptoms become apparent when the ratio Mn to Fe concentration
falls below 18 (Lee, 1972). Application of Fe as FeEDTA to the foliage of E. gummifera seedlings exhibiting
Mn toxicity resulted in recovery of chlorotic leaves (Winterhalder, 1963).
The “crinkle-leaf” symptoms are similar to Ca deficiency symptoms, are associated with a decrease in Ca
translocation to leaves and are often classified as an induced Ca deficiency (Horst and Marschner, 1978).
However, leaves displaying crinkle-leaf symptoms in G. max while having lower Ca concentrations than in
control plants were still within the normal Ca range (Wu, 1994).
Negative relationships have been determined between Mn in the external medium and tissue Mg concentration in
Pinus radiata needles but not in roots (Safford, 1975). This suggests that Mn toxicity has an inhibitory affect on
Mg translocation within the plant and could suggest that high Mn induced Mg deficiency, especially at low
external Mg supply.
4.4.3 Zinc
4.4.3.1 Toxicity symptoms
The first symptom to present itself in most species exhibiting Zn toxicity is a general chlorosis of the younger
leaves (Harmens et al., 1993; Ren et al., 1993; Fontes and Cox, 1995). Depending on the degree of toxicity this
chlorosis can progress to reddening due to anthocyanin production in younger leaves (Harmens et al., 1993;
Fontes and Cox, 1995; Lee et al., 1996a). Plants exhibiting Zn toxicity have smaller leaves than control plants
(Ren et al., 1993). Glycine max plants normally have horizontally orientated unifoliate leaves. However, Zn
stressed plants exhibit vertically oriented leaves (Fontes and Cox, 1995). Brown spots become apparent on the
leaves of some species (Fontes and Cox, 1995). In severe cases plants may exhibit necrotic lesions on leaves and
eventually entire leaf death (Harmens et al., 1993). In roots, Zn toxicity is apparent as a reduction in the growth
of the main root, fewer and shorter lateral roots and a yellowing of roots (Ren et al., 1993).
23
4.4.3.2 Symptoms of induced deficiency
Many studies have shown that the chlorosis which presents itself as a symptom of Zn toxicity is coupled with a
lowering of Fe uptake as Zn concentration in the external medium increases (e.g. Wallace and Abou-Zamzam,
1989; Ren et al., 1993). Painting of leaves with Fe solutions, such as ammonium ferric sulphate, leads to a
regreening of chlorotic leaves (Chaney, 1993; O'Sullivan et al., 1997). As such, symptoms of chlorosis are
associated with an induced Fe deficiency.
Care must be taken that experiments describing Zn-induced Fe deficiency have rigorously tested this hypothesis.
For example, two papers by Fontes and Cox (1998b; 1998a) suggest Zn induced Fe deficiency at high Zn and
low Fe as this was corrected when they added extra Fe to the solution. Speciation analysis of the solutions using
GEOCHEM (Sposito and Mattigod, 1980) suggested other more likely scenarios. The Fe was added to the
nutrient solution as FeEDTA and most of the Zn in the high Fe treatments would be complexed with the EDTA
resulting in free Zn concentrations that are not toxic. As well, at low Fe concentrations most of the Fe would be
present as solid Fe-phosphate suggesting that the Fe deficiencies may be direct Fe deficiency without any
confounding effects of Zn toxicity.
In the presence of high Zn concentrations, Zn sensitive grasses have been shown to develop P deficiency
symptoms although P tissue concentrations appeared normal (Plenderleith, 1984). Similarly, across a wide
variety of crop and pasture species grown under excess Zn, symptoms were similar to those for P deficiency but
P tissue concentrations did not decrease (Boawn and Rasmussen, 1971). These findings suggest that Zn toxicity
may result in P deficiency caused not by effects on P uptake but by effects on P metabolism.
4.5 Effects of toxicity on physiology
The toxic action of metals is seen at a macro-scale on growth reduction and foliar symptoms. At a smaller scale,
the effects of metals can be seen anatomically as cellular symptoms. Symptoms, both macro and cellular, and
growth reduction are side effects of the direct mode of action. The direct mode of action of a metal is on plant
metabolism. Increase in the concentration of any metabolite can demonstrate either increased production or
decreased utilisation by the reactions for which it is a substrate or product (Burke et al., 1990).
4.5.1 Photosynthesis
At elevated Cu concentration, where root symptoms were apparent but not growth reductions, total chlorophyll
contents and chlorophyll a to b ratios were reduced but there was no effect on net photosynthesis (Rousos et al.,
1989). At higher external Cu concentrations where growth was depressed, lower chlorophyll contents
(Ouzounidou et al., 1992; Luna et al., 1994; Ouzounidou et al., 1994), a reduced photosynthetic capacity,
including an inhibitory effect on photosytem II (Arellano et al., 1995; Baron et al., 1995), and an increase in the
break down of carotenoid (Luna et al., 1994) occurred. It has been demonstrated that the toxic action is
substitution of Cu for Mg in chlorophyll molecules, thus reducing photosynthesis (Kupper et al., 1996). A side
effect of Cu inhibiting photosynthesis is an increase in the production of free radicals and therefore an increase
in rate of leaf senescence due to oxidative damage (Luna et al., 1994). However, in some species an increase in
photosynthetic pigments occurs because of a Cu induced reduction in CO2 fixation and, as such, photosynthesis
does not decrease, at least initially (Romeu-Moreno and Mas, 1999).
Plants have been shown to react to excess Mn with a drop in photosynthetic rate (Macfie and Taylor, 1992;
Macfie et al., 1994). This lowering of photosynthesis occurred as a result of decreases in chlorophyll and the
photosynthesis per unit chlorophyll in a sensitive cultivar and only reducing the chlorophyll content in a tolerant
cultivar of T. aestivum (Moroni et al., 1991; Macfie and Taylor, 1992). Manganese toxicity has also been
associated with swollen chloroplasts in G. max (Wu, 1994).
Toxic concentrations of Zn in P. vulgaris resulted in inhibition of photosystems I and II and thus a decrease in
photosynthesis (van Assche and Clijsters, 1986). It has been demonstrated that the mechanism of action is the
displacement of Mg by Zn at the water splitting site in photosystem II (van Assche and Clijsters, 1986; Kupper
et al., 1996). In Spinacia oleracea (spinach), excessive Zn supply was found to greatly reduce ATP synthesis
and activity in chloroplasts (Teige et al., 1990). After 50 minutes treatment in 1mM Zn the uncoupled electron
transport rate was reduced by 26%. Thylakoids treated with 1 to 2mM Zn for 30 minutes exhibited inhibition of
the water splitting site of photosystem II (Teige et al., 1990). Inhibition caused by Zn could be restored by
benzidine which is an electron donor for photosystem II (Teige et al., 1990). However, after 60 minutes of
exposure irreversible membrane damage occurred which could not be restored by benzidine. Teige et al. (1990)
suggested that the primary toxic action of Zn is the inhibition of ATP synthesis and therefore energy metabolism
in plants.
24
4.5.2 Transpiration and water budgets
Little research has been undertaken into the effects of Cu or Zn toxicity on transpiration and water budgets
within plants. However, at low Cu toxicity in Brassica. oleracea var. capitata (cabbage) where root symptoms
are expressed but growth has not decreased no effect on transpiration was apparent (Rousos et al., 1989).
Transpiration in H. vulgare has been shown to be only slightly inhibited at high Zn concentrations (Brune et al.,
1994). However, wilting, a qualitative sign of effects on water budgets, has been observed in plants exposed to
high Zn concentration (Ye et al., 1998). Hence, it appears as though excess Cu and Zn have little influence on
plant transpiration and water budgets.
In P. vulgaris, the stomatal and cuticular transpiration per unit leaf area increased under high Mn (Horst and
Marschner, 1978). Horst and Marschner (1978) suggest that the higher transpiration rate is probably caused by a
loss of the metabolically controlled stomatal movements caused by Mn toxicity. In contrast, Suresh et al. (1987)
found that stomatal conductance and transpiration decreased as soil Mn increased for G. max. The effect was
more pronounced in the less tolerant cultivar (Suresh et al., 1987) suggesting that lowering of transpiration was
utilised in a tolerance response. In another study on P. vulgaris, decreased stomatal conductance and
transpiration rates were only apparent in tissue with advanced chlorosis, with leaves only exhibiting brown
spotting not affected (Gonzalez and Lynch, 1997).
4.5.3 Enzymes and cell metabolism
Copper toxicity has a significant effect on enzyme production and metabolism. Excess Cu has been shown to
inhibit ATPase activity in the plasma membrane of Z. mays roots (Kennedy and Gonsalves, 1989). However, the
authors suggested that this was an indirect effect of Cu toxicity resulting from the leakage of K ions. Excess Cu
inhibited acid phosphatase activity in Deschampsia cespitosa (Cox and Hutchinson, 1980). Copper toxicity has
been associated with an increase in antioxidative enzymes as a result of Cu meditated oxidative damage (Luna et
al., 1994; Weckx and Clijsters, 1996; Savoure et al., 1999). The contents of amino acid-N and protein-N in the
shoots, but not the roots, of Silene vulgaris decrease with increasing Cu concentration (Weber et al., 1991).
Hence, Cu had a negative effect on the metabolism of N, amino acids and proteins within the shoots of plants.
Little research has been conducted into the effects of Mn toxicity on the metabolism of enzymes, proteins and
amino acids. Peroxidase activity was found to increase in O. sativa plants with increasing external and internal
Mn concentrations (Horiguchi, 1988). This author considered it likely that the high peroxidase activity was
associated with the mechanism of necrotic browning in Mn affected plants.
Excess Zn has been shown to stimulate the production of a range of enzymes in P. vulgaris (van Assche et al.,
1988). Van Assche et al. (1988) suggested that this might be a compensation by the cell for the inhibition of
physiological activity caused by high Zn, such as the inhibition of chloroplast NADPH production. Free radical
generation was accelerated in plants exposed to excess Zn (Prasad et al., 1999). In an attempt to counteract the
toxic effects of high O, enzyme-specific activities of superoxide dismutase, catalase, guiacol peroxidase,
ascorbate peroxidase, monodehydroascorbate reductase, dehydroascorbate reductase and glutathione redeuctase
were found to increase (Prasad et al., 1999). In O. sativa, high external Zn concentration was demonstrated to
increase peroxidase, auxin oxidase and ascorbic acid oxidase whereas the activity of catalase IAA oxidase, αamylase, ATPase and phytase was inhibited (Nag et al., 1984). Acid phosphatase production was inhibited in
both tolerant and non-tolerant clones of Anthoxanthum odoratum, but more so in the non-tolerant clone, after
exposure to elevated Zn (Cox and Hutchinson, 1980).
Zea mays plasma membranes increased activation of ATPase up to 3 mM Zn (Kennedy and Gonsalves, 1989).
The authors suggested that Zn was able to substitute, although imperfectly, for Mg as a substrate for the enzyme.
High external concentrations of Zn have been found to inhibit RuBP carboxylase activity in P. vulgaris however
RuBP oxygenase was not affected (van Assche and Clijsters, 1986). These results also suggested that Zn
partially substituted for Mg in the RuBisCo-complex (van Assche and Clijsters, 1986).
Gibberellic acid has been shown to exhibit a protective effect against Zn toxicity in O. sativa seedlings (Nag et
al., 1984). It was suggested that this may be associated with the action of gibberellic acid on the enhancement of
synthesis of enzyme proteins.
Zinc toxicity has been found to result in a decrease in amino acid accumulation in Panax quinquefolium
(American ginseng) roots (Ren et al., 1993). In contrast, in D. cespitosa roots, toxic concentrations of Zn were
shown to cause accumulation of asparagine and proline while other amino acids were not significantly changed
25
(Smirnoff and Stewart, 1987). Not enough research appears to have been conducted to derive any conclusions
on the effect of Zn toxicity on amino acid concentrations in plant tissue.
From 12 to at least 96 hours after the initiation of high Zn exposure, roots of a tolerant Festuca rubra cultivar
had higher protein contents in their mitotic cells compared to controls, whereas a sensitive cultivar had lower
protein levels than its control (Davies et al., 1991b). The increase in the tolerant cultivar may have been due to
the production of 'stress proteins' which protected the plant. In contrast, the decrease in the sensitive cultivar is
likely the result of a toxic effect on cell metabolism (Davies et al., 1991b).
4.5.4 Cell functioning
In vitro studies with cells of H. annuus roots have shown that, as the Cu supply increased, the mitotic index
decreased, and mitotic abnormalities increased (Chakravarty and Srivastava, 1992). In comparison, Pinus plants
grown in vivo had no decrease in mitotic index in root tip cells on exposure to elevated Cu a decrease in root
elongation and growth (Arduini et al., 1995). The authors suggested that the lack of effect on mitotic index was
due to the protective effect of the root cap on the meristematic cells. As well as decreasing cell division, Cu
toxicity has been shown to reduce cell elongation (Romeu-Moreno and Mas, 1999).
Excess Cu results in a reduction in plasma membrane integrity in plant roots and it is thought that this is the
mechanism by which Cu toxicity retards root growth (de Vos et al., 1991; Luna et al., 1994; Arduini et al.,
1995). This has been associated with the leaking of K from Z. mays roots (Kennedy and Gonsalves, 1989).
Under conditions of Mn toxicity, leaf cell volume decreased before cell number (Terry et al., 1975). However,
the authors stress that this does not mean cell division is less sensitive to Mn toxicity than cell expansion as leaf
cell numbers are determined early in the life of the leaf when leaf Mn concentration was probably at its lowest.
Leaf tissues from areas exhibiting crinkle-leaf symptoms have been shown to have jagged epidermis and
disintegration of cells (Wu, 1994). Necrosis was also associated with collapse of mesophyll and epidermal cells
in leaves from plants exposed to excess Mn (Dienelt and Lawson, 1991). The concentration of extractable leaf
callose, an indication of injury, correlates highly with Mn toxicity symptoms (Wissemeier et al., 1992).
High external Zn concentration inhibited stem cell elongation (Aidid and Okamoto, 1992). However the
membrane potential across the xylem-symplast interface was not affected, suggesting that this is not the active
site of Zn toxicity (Aidid and Okamoto, 1992). As Zn supply increased, H. annuus cells had an increase in
mitotic abnormalities and a decrease in mitotic index in in vitro studies (Chakravarty and Srivastava, 1992).
4.6 Mechanisms of tolerance
There is much contention in the literature over the possible mechanisms of metal tolerance. This could indicate a
general lack of understanding of metal toxicity issues but it is just as likely to reflect the complex nature of
higher plant responses to metal toxicity. It is quite likely that different species may have evolved different
mechanisms to tolerate excess metals and that even within the one plant species more than one mechanism could
be in operation. Plants have both constitutive (utilised by sensitive and tolerant phenotypes) and adaptive
(utilised by tolerant phenotypes only) mechanisms to withstand excess metals (Meharg, 1994). This review is
primarily considering adaptive tolerance mechanisms as it is these that shed light on why some phenotypes are
able to withstand higher metal exposure than others. There are a number of strategies that plants could possibly
employ to combat high external metal concentrations. These can be classified in two main categories, i.e. firstly,
restriction of uptake or transport and secondly, internal tolerance mechanisms.
4.6.1 Restriction of uptake or transport
4.6.1.1 Exclusion from the plant
One mechanism of preventing or lessening the toxicity effects of metals is preventing excess metals from
entering the plant. There are thought to be two main ways in which a plant could do this, either by precipitating
or by complexing metals in the root environment.
Plants could precipitate metals by increasing the pH of the rhizosphere or by excreting anions such as phosphate.
While there is a large body of circumstantial evidence suggesting a pH mediated tolerance mechanism, it is
becoming apparent that pH may play a minor role but it is unlikely to be a major tolerance mechanism (Taylor,
1991). Studies where the form of N supply is manipulated, and therefore also plant induced pH, contradict the
pH hypothesis. Plants grown in high NO3- solutions do not show greater tolerance to Zn than plants grown in
26
NH4+ (Smirnoff and Stewart, 1987). Although care should be taken before dismissing this hypothesis with
arguments utilising the pH of the bulk nutrient solution. Minimal changes in bulk solution pH may actually be
significant in the rhizosphere soil solution . Species of Lotus differing in Al tolerance have been shown to have
no difference in solution, rhizosphere, or root surface pH (Blamey et al., 1990b). Research with an Al tolerant
mutant of Arabidopsis thaliana (alr-104) has shown that, upon exposure to Al, the mutant increased the solution
pH by 0.2 (4.3 to 4.5) of a unit while the wild type solution pH remained the same (Degenhardt et al., 1998).
Further experiments showed that increasing solution pH by 0.1 to 0.2 units was enough to nearly double the
growth rate of both the mutant and the wild type under Al exposure and that buffering the solution pH prevented
the Al tolerance advantage of the mutant. This research by Degenhardt et al. (1998) appears to be the first
thorough example of metal tolerance gained by increases in rhizosphere pH.
Root exudation of P in Z. mays has been detected at Al3+ activities where the tolerant cultivar expressed no
toxicity symptoms and the sensitive cultivar showed symptoms (Pellet et al., 1995). This suggests that the
tolerant cultivar may have utilised exclusion of the metal via precipitation as a tolerance mechanism.
Iron plaque formation on roots has been suggested as a mechanism for excluding toxic metal, particularly in O.
sativa (Ye et al., 1998). It appears, though, that Fe plaque does not protect against Zn, Cd or Pb toxicity per se
and may act instead by improving the Fe status of plants (Ye et al., 1998).
The theory behind plants excreting chelators, is that the chelators could complex heavy metals in the vicinity of
the root thus making them unavailable to the plant and lessening the experienced toxicity. Organic acids have
been suggested as likely chelating substances. Tolerant T. aestivum genotypes have been found to produce more
malic acid than sensitive genotypes upon exposure to Al (Delhaize et al., 1993; Huang et al., 1996). In the same
paper, malic acid added to solution cultures was able to protect T. aestivum from Al toxicity. Work by Pellet et
al. (1995) has demonstrated in Z. mays at elevated Al that citric acid exudation was increased only by tolerant
cultivars. A tolerant cultivar of P. vulgaris has been found to excrete larger amounts of citric acid when grown
in the presence of Al and when compared to a sensitive cultivar (Miyasaka et al., 1991). Fagopyrum esculentum
(buckwheat), an Al tolerant species, has been shown to excrete high quantities of oxalic acid when exposed to
excess Al and sufficient P (Zheng et al., 1998). This indicated that oxalic acid secretion was a true tolerance
mechanism and not an induced deficiency response. Aluminium tolerant mutants (alr-108, alr-128, alr-131 and
alr-139) of A. thaliana were found to take up less Al and produce more citrate and malate than the wild type
(Larsen et al., 1998). However, the production of organic acids was not Al mediated. In contrast, no difference
was found between the complexed Al (indicating organic acids) in the nutrient solutions of two species of Lotus
differing in their Al tolerance, though it should be noted that the solution were not sterile (Blamey et al., 1990b).
Much research has found direct evidence of metal exclusion from the plant. The most likely widespread
mechanism for exclusion appears to be the exudation of organic acids. However, much of the work has been
conducted with Al and so more work with other metals is required to test the widespread validity of this
mechanism. It must also be remembered that while excluding excess metals from the plant holds merit, some
tolerant plants actually have a higher uptake of metals when compared to sensitive plants (Smirnoff and Stewart,
1987; Harmens et al., 1993) so other mechanisms must also exist.
4.6.1.2 Cellular exclusion
A large fraction of metals in plant roots are found in the apoplastic free space. At equal external Al
concentrations a sensitive T. aestivum cultivar had more symplastic Al than the tolerant cultivar suggesting an
exclusion mechanism (Tice et al., 1992). However work with grasses has shown that the total amount of metal
present in apoplast, while increasing with external Zn supply, was similar in tolerant and sensitive plants
(Brookes et al., 1981).
4.6.1.3 Complexation at the cell wall-plasma membrane interface
Significant proportions of metal accumulate at the cell wall-plasma membrane interface and it has been
hypothesised that this could be the site of metal tolerance. For example, 60% of Cu in the roots of both Lolium
multiflorum (Italian ryegrass) and T. pratense was bound by the cell wall and plasma membrane (Iwasaki et al.,
1990), although no measures were made of plant tolerance. Minuartia verna ssp. hercynica growing on heavy
metal contaminated medieval mine dumps has been found to have high concentrations of Fe, Cu, Zn and Pb
associated with Si contained in the cell walls (Neumann et al., 1997). In comparison, no accumulation of heavy
metal was detected in the cytoplasm suggesting strong use of exclusion by the metal adapted subspecies.
However, the use of glutaraldehyde fixation techniques has been shown to result in the loss of up to 70% of the
27
metal present in the plant (Neumann et al., 1997). Thus the metal distribution pattern may largely be an artifact
of the fixation process.
Plant cation exchange capacity (CEC) is largely determined by the exchange sites in cell walls (Horst and
Marschner, 1978). Sensitive cultivars of T. aestivum have been shown to have much lower cell wall CECs than
tolerant cultivars (Masion and Bertsch, 1997). It has been suggested that tolerant cultivars use the high CEC to
complex metals at the cell wall and prevent entry to the cell. Horst and Marschner (1978) found that as Mn
uptake and toxicity symptoms increased in P. vulgaris plants the leaf CEC decreased. This suggests that a low
CEC is a symptom of toxicity. Horst and Marschner (1978) hypothesised that low CEC inhibits Ca translocation
and thus the protective effect of Ca. In contrast, low root CECs were found to correspond to Al tolerance in
Lotus species (Blamey et al., 1990a), although it should be noted that this was whole root rather than cell wall
CEC per se. The hypothesis here being that for cell wall pectin to continue playing a protective role at high
solution Al levels it needs a low CEC, thus low methylation and low levels of pectin precipitation by Al (Blamey
et al., 1990a). However, it is recognised that different species exposed to different metals have varying CEC
responses. In P. vulgaris, P. sativum, L. esculentum, H. annuus and Avena sativa (oats), Zn, Ni and Co toxicity
increased root CEC (Crooke, 1958). In contrast Mn reduced root CEC and Cu caused an increase in A. sativa
and H. annuus but a reduction in the other crops tested (Crooke, 1958)
Clearly, the role of the cell wall in tolerance is complicated, with a number of different mechanisms possible,
some of which can act in more than one, often contradictory way.
4.6.1.4 Active efflux
Whilst a possible mechanism of metal tolerance, little research has been conducted into active efflux. It is likely
that this lack of research is because of the difficulty of distinguishing between lack of influx and efflux. Work
with Lupinus albus suggests Cd excretion by roots as a possible tolerance mechanism (Costa and Morel, 1993).
While not a root-based mechanism, M. verna ssp. hercynica has been shown to excrete heavy metals through
hydathodes (Neumann et al., 1997). Shedding of metal loaded older leaves is another possible mechanism of
active efflux. Loading older leaves with metals and the shedding of older leaves has been noted in many species
under metal induced stress (Luna et al., 1994; O'Sullivan et al., 1997) although it has rarely been connected
directly to a tolerance mechanism.
4.6.1.5 Uniformity of distribution
By having a nonuniform distribution and translocation of metals within the plant, plants may be able to minimise
the effects of excess metals. The most common form of this is the much higher amounts of metals found in plant
roots as compared to shoots (Horst and Marschner, 1978; Brown and Wilkins, 1985a; Qureshi et al., 1985;
Wheeler and Power, 1995; Zhang et al., 1998). Presumably, this acts to exclude metals from more sensitive
metabolism in the shoots. Root/ shoot grafting experiments with G. max cultivars has shown that, in this species,
Zn tolerance is conferred by shoot genotype (White et al., 1979). However, while shoot concentrations were
lower in some tolerant phenotypes (Brown and Wilkins, 1985a; Ouzounidou et al., 1994), other work showed
shoot metal levels in tolerant phenotypes to be higher than in sensitive ones (Qureshi et al., 1985). This suggests
that reduced translocation to the shoot is unlikely to be a widespread tolerance mechanism.
Tolerant V. unguiculata genotypes have a relatively uniform distribution of Mn throughout the leaf (Horst,
1983). In contrast, sensitive genotypes accumulate Mn in localised areas in the leaf which are seen as dark
brown spots caused by precipitates of Mn oxides (Horst, 1983). Similar patterns have been recorded at a cellular
level in Z. mays under Cu stress. In Z. mays roots some cells in a tissue are damaged by excess Cu and others
continue to function normally (Ouzounidou et al., 1995). It thus appears that some cells accumulate Cu and die
while allowing others to maintain their functioning (Ouzounidou et al., 1995). This non-uniform response of
cells to Cu would appear to be a mechanism to overcome Cu toxicity, however whether or not it was utilised
more by especially tolerant phenotypes is not known.
4.6.2 Compartmentation and complexing within the cell
4.6.2.1 Compartmentation within vacuoles
Metal tolerance could be achieved if metals were sequestered away in places within the cell where the metals
cannot react with metabolically active cellular substances. Compartmentation in the vacuole is regularly put
forward as the most probable site. It has been demonstrated that grasses can actively pump Zn into vacuoles
with the more tolerant clones being able to continue the process at higher external Zn levels than sensitive clones
28
(Brookes et al., 1981). In H. vulgare leaves, the increase in cellular Zn with increasing exposure to external Zn
was fully accounted for by an increase in vacuolar Zn with the cytoplasm exhibiting perfect homeostasis (Brune
et al., 1994). Upon exposure to Al, a tolerant Z. mays variety produced vacuoles in the root apex cells, however,
this was not tested against a sensitive variety (Vazquez et al., 1999). These vacuoles were rich in Al which was
associated with a substance with composition similar to phytate and also with Si. Excess Zn increased the
vacuolation of both a sensitive and tolerant cultivar of F. rubra but the increase and the amount of Zn
sequestered was greater in the tolerant cultivar (Davies et al., 1991a). Other species where the vacuole has been
suggested as the site of sequestration include D. caespitosa grown with excess Zn (van Steveninck et al., 1987)
and Nicotiana tabacum (tobacco) grown with excess Cd (Vogeli-Lange and Wagner, 1990). Copper toxicity has
been shown to increase vacuole numbers in meristems of Z. mays roots (Doncheva, 1998). Computer simulation
studies with Zn (Wang et al., 1992) and Cd (Wang et al., 1991) have shown that the vacuole is the probable site
of metal sequestration and detoxification in N. tabacum. Due to the wide range of testing conditions used in the
computer simulations it is possible that the results would transfer to other species. Only a few papers report
negative findings on the importance of vacuoles in metal tolerance (e.g.Tice et al., 1992). Evidence for plant
vacuoles as the site of metal sequestration appears to be quite conclusive although questions remain as to the
pervasiveness of this mechanism with metals other than Zn and Cd, and in dicots.
Research with a highly tolerant Mn species, H. annuus, has shown accumulation and excretion excess Mn in leaf
trichomes (Blamey et al., 1986). Subsequent work with Vicia faba (faba bean) showed that it expresses
metallothionein genes in the trichomes suggesting the possibility of a mechanism combining metallothionein
production with sequestering in the trichomes(Foley and Singh, 1994).
4.6.2.2 Complexing by metallothioneins
Metallothioneins are a group of low molecular mass, cysteine rich, metal-binding proteins (Robinson and
Jackson, 1986; Tomsett and Thurman, 1988). It has been suggested that they function in the regulation of
essential metals and in the detoxification of all metals (Steffens, 1990). Metallothioneins were originally found
in animals and this prompted the search for them in plants as a detoxification mechanism (Tomsett and Thurman,
1988). However it has since been shown that the case for metallothioneins is not as strong as in animals
(Tomsett and Thurman, 1988; Steffens, 1990), and it is now often assumed that metallothioneins have no
function in metal tolerance rather than being substantiated by hard evidence (e.g.Zenk, 1996)
A cysteine rich, low molecular weight Cu-binding protein found in Agrostis gigantea was given as evidence for
the role of metallothioneins in plant metal tolerance (Rauser, 1984). However, the plants were not compared
with a control nor were they compared with a sensitive cultivar. A tolerant clone of F. rubra produced more
protein in mitotic cells when exposed to excess Zn than control plants or a sensitive cultivar (Davies et al.,
1991b), however the nature of the protein and its Zn binding capacity were not explored. Production of
metallothioneins in Arabidopsis ecotypes was found to be correlated with Cu tolerance, and in dose response
experiments Arabidopsis ecotypes were found to have saturation levels of metallothioneins at the onset of
symptoms (Murphy and Taiz, 1995).
Tukendorf et al. (1984) found that tolerant S. oleracea plants produced greater amounts of proteins when
exposed to excess Cu, and that most of the Cu in the plant was bound to proteins when compared to control
plants. These proteins had a lower cysteine content than metallothioneins but were suggested as a mechanism
similar to metallothioneins in reducing Cu toxicity. However, no comparison was made with sensitive cultivars,
so the effect may be just as strong or stronger in sensitive plants, which would imply a stress rather than a
tolerance response.
Sulphur (i.e. thiol) rich proteins has been used as an indirect measure of metallotheioneins. Deschampsia
cespitosa has been shown to not utilise S-rich proteins in three ways. Firstly, in S deficiency situations D.
cespitosa produces less Cu induced S-rich proteins (Schultz and Hutchinson, 1988). However, tolerant clones
exhibited the same amount of tolerance to excess Cu while non-tolerant clones become even more sensitive
under S-deficiency. Secondly in D. cespitosa the production of S-rich proteins was not found to increase with
time of exposure to high Cu. Thirdly, tolerant clone of D. cespitosa produced less S-rich proteins than nontolerant clones. This evidence suggests that metalloproteins could possibly be a toxicity stress response rather
than a tolerance response. However, the results of Schultz and Hutchinison (1988) must be viewed with some
caution, as thiol-rich protein is an indirect and unreliable estimate of metallothioneins (Murphy and Taiz, 1995).
Metalothioneins or metalotheionein-like proteins were not found in tolerant or sensitive leaves in P. sativum
plants exposed to excess Cu, however a Cu containing peptide in tolerant but not sensitive cultivars was
identified and may play a role in tolerance (Palma et al., 1990).
29
The latest research into metallothioneins and tolerance has been with gene expression. Metallothionein genes
have been found in V. faba, although expression of the gene was not increased upon exposure to Cu, Zn or Cd
(Foley and Singh, 1994). Research with L. esculentum has shown that it has at least five metallothionein coding
genes (Giritch et al., 1998). Different genes appeared to be expressed in different organs with Cu-induced genes
being expressd mainly in the roots and Zn-induced genes being mainly expressed in leaves upon exposure to
these metals. This suggests that some studies which did not find a relationship between metal exposure and
metallotheionein production may have been considering the wrong plant organ.
While metallothioneins do exist in the plant kingdom, their widespread role in adaptive tolerance to metals is
contentious. It has been suggested that metallothioneins may have a role in trace metal metabolism and cell
homeostasis rather than metal tolerance per se. (Robinson and Jackson, 1986; Foley and Singh, 1994; Giritch et
al., 1998).
4.6.2.3 Complexing by phytochelatins
Phytochelatins are cysteine-rich non-protein metal-binding peptides produced by plants (Schat and Kalif, 1992;
Zenk, 1996). It has been suggested that they may act in a similar way to that proposed for metallothioneins in
the plant tolerance to metals.
Zea mays has been found to have phytochelatins which bind to excess Cu and Cd (Galli et al., 1996). However
this was not tested against a control or sensitive cultivar. Phytochelatin synthesis was initiated in Rubia
tinctorum root cultures upon exposure to a number of metals including Zn, Cu and Cd, and no phytochelatin
production was observed in controls plants (Maitani et al., 1996). Exposure of N. tabacum seedlings to Cd
resulted in the production of phytochelatins (Vogeli-Lange and Wagner, 1990). Computer simulation studies
with N. tabacum suggested that phytochelatins could have a major role in tolerance to Cd assuming they were
produced upon exposure to Cd (Wang et al., 1991). Positive research for a role of phytochelatins in metal
tolerance seems to assume that the presence of phytochelatins implies they are there in restorative role rather
than testing their production in sensitive versus tolerant phenotypes.
Indirect evidence for the role of phytochelatins in plant tolerance to metals comes from experiments with
buthionine sulphoximone (BSO). As BSO is an inhibitor of a precursor to phytochelatin synthesis, it can be used
to prevent phytochelatin production (e.g.Ruegsegger et al., 1990; Maitani et al., 1996). Addition of BSO to
nutrient solutions has resulted in increased sensitivity of Betula pendula (silver birch) to Cd (Gussarsson et al.,
1996), Silene cucubalus to Cu and Cd (de Vos et al., 1992) and N. tabacum cells cultures to Cu and Zn (Reese
and Wagner, 1987) suggesting a role for phytochelatins in plant metal tolerance. In N. tabacum cell cultures
BSO plus Cd has been demonstrated to reduce cell growth even at Cd levels which did not normally cause
growth reductions, whereas BSO with no Cd had no effect on growth (Reese and Wagner, 1987).
Research with tolerant and non-tolerant strains of S. vulgaris exposed to a range of Cu concentrations have
shown that phytochelatins are not involved in the tolerance process in this species for three reasons (Schat and
Kalif, 1992). Firstly, the phytochelatin to Cu ratio is much lower in the roots of tolerant plants, meaning that
proportionally more Cu is bound in non-tolerant plants. Secondly, the solution concentration of Cu required to
increase total non-protein sulphydryl and the solution concentration at maximum sulphydryl production are
lowest in tolerant plants. Thirdly, non-tolerant plants have a higher non-protein sulphydryl content on a whole
root basis. Phytochelatins have also been shown to not play a role in Cd tolerance for S. vulgaris (de Knecht et
al., 1994).
It has been suggested that phytochelatins play a constitutive role in plant metal tolerance (Meharg, 1994; Zenk,
1996). However, evidence provided by Zenk (1996) for this role is not conclusive and could just as easily
indicate a stress response (e.g. production of phytochelatins upon exposure to metals). Meharg (1994) using the
research of de Vos et al. (1992), and suggested that tolerant S. vulgaris may use phytochelatins as a constitutive
mechanism and have as their adaptive mechanism a process which prevents access of metals to the cytoplasm.
Research by Keltjens and van Beusichem (1998) has suggested that phytochelatin production may be initiated by
root damage irrespective of whether this is caused by metals. It has also been shown that phytochelatins have a
role in metal transport (Vogeli-Lange and Wagner, 1990; Salt and Rauser, 1995), so that any detoxifying
capabilities may actually be secondary or part of a more complex mechanism. The role of phtyochelatins in
metal tolerance reminas poorly understood and requires more thorough research.
30
4.6.2.4 Complexing by organic acids
Organic acids act as metabolic intermediates in the formation of ATP from carbohydrates in N metabolism and
in ionic balance. Hence, metabolic abnormalities in any of these processes would be reflected by changes in the
concentrations of the intermediate organic acids (Burke et al., 1990). Therefore, an increase in organic acids
with increasing supply of metals could imply a detoxification mechanism or, conversely, disruption of
metabolism resulting in the production of organic acids as a stress response to excess metal.
Organic acids within cells could act to detoxify metals by complexing them and making them unavailable to the
plant. There is contradictory evidence with research into organic acids within plant cells as a tolerance
mechanism. Researchers have found that organic acid levels can increase in both tolerant and sensitive cultivars
depending on the species and metal studied. Manganese tolerant cultivars of T. aestivum did not increase their
production of malic, citric or aconitic acid when exposed to high solution Mn concentrations while
concentrations of the organic acids slightly increased in the sensitive cultivars (Burke et al., 1990). In D.
caespitosa malic acid production was increased proportionately more in tolerant clones exposed to excess Zn
(Brookes et al., 1981). Copper and Zn resistant Nicotiana plumbaginifolia cell cultures produced 3 to 12 times
the citrate and malate of non-tolerant cells and slight increases of the non-tolerant cells for succinate and
fumarate when exposed to high Zn or Cu (Kishinami and Widholm, 1987). In Z. mays plants, the relative
production of different organic acids was found to vary with external Al concentration, suggesting that the metal
concentration is important for organic acid production (Pintro et al., 1997). At 3.4 µM Al activity t-acontic,
citric, formic, malic, oxalic and succinic acids were significantly higher in a tolerant Z. mays cultivar while
quinic acid concentration in a sensitive cultivar was elevated (Pintro et al., 1997). By 10.3 µM Al activity, only
t-aconitic acid remained higher in the tolerant cultivar. This suggests that some organic acids may play a role in
metal tolerance but that by 10.3 µM most of the organic acid based tolerance mechanisms in the tolerant cultivar
were no longer in operation. A Zn and Cd tolerant cultivar of F. rubra was found to produce more malic acid on
exposure to excess Zn compared with a sensitive cultivar where malic acid production changed little. However,
when exposed to excess Cd, both cultivars reduced production of malic acid, suggesting that Zn and Cd tolerance
may operate by different mechanisms (Harrington et al., 1996).
It has been shown by computer simulation studies that citrate is likely to be a major complexing agent of Zn in
N. tabacum, and oxalate is likely to be a major complexing agent of Zn in high oxalate producing plants (Wang
et al., 1992). However computer simulations with Cd suggest that organic acids are only likely to play a minor
detoxifying role (Wang et al., 1991).
Detoxification of metals by organic acids appears to hold merit. However, further research is needed to quantify
the organic acids being produced as a toxicity versus tolerance response and to test the prevalence of this
mechanisms in different species with various metals.
4.6.2.5 Complexing with inorganic and organic ligands
Phytate has been suggested as a possible complexing agent for metals. Phytate tends to be associated with metal
deposits in many plants including in Al tolerant Z. mays vacuoles (Vazquez et al., 1999). Zinc-phytate globules
appeared to be qualitatively more frequent in tolerant ecotypes of D. caespitosa roots than in sensitive ecotypes
(van Steveninck et al., 1987). Phosphorus is an indirect measure of phytate and has been found associated with
metal deposits in plants including Zn in the vacuoles of Minuartia verna (Davies et al., 1991a). Deposits of ZnP have been found in G. max, L. sativa and Z. mays but not in H. annuus, P. sativum and Lolium sp. (ryegrass)
when exposed to high Zn (van Steveninck et al., 1994). The same authors found no evidence for Cd-P deposits
in any of the species studied when exposed to high Cd. However, the role of phytate in tolerance does not
appear to have been directly studied and so it is unclear if metal phtytae complexes are a tolerance or stress
response.
Computer simulations suggest that inorganic ligands would not play an important role in Zn detoxification in N.
tabacum (Wang et al., 1992). The same techniques for Cd only suggest a role for inorganic ligands (Cdsulphate) at very high Cd (Wang et al., 1991). While these studies suggest that inorganic ligands may act to
detoxify metals, this straegy has not been widely studied.
4.6.2.6 Alterations of cellular metabolism
There is a possibility that other metabolic changes, besides metal-complexation, may have a role in plant metal
tolerance. This could occur by two pathways. Firstly, metal-sensitive metabolic processes could be avoided by
the activation of alternative pathways. Secondly, the sensitivity of enzyme activity to metals could be
31
counteracted by increasing the production of enzymes. However, neither possibility has been substantiated by
direct evidence (Brookes et al., 1981; Verkleij and Schat, 1990; Taylor, 1991).
Circumstantial evidence includes that of Ma et al. (1998) who have shown that, while oxalic acid does not
increase in F. esculentum cell sap upon exposure to 50 µM Al, it is responsible for complexing the majority of
the Al in the cell sap. Fagopyrum esculentum is more tolerant of Al than many other species (Zheng et al.,
1998) and so the high natural concentrations of oxalic acid in F. esculentum cells may indicate a metabolism
evolved to excess Al. Van Assche et al. (1988) have demonstrated that the capacity of numerous enzymes
increases with leaf Zn and Cd content in P. vulgaris. However, this could just as easily be a stress response.
Acid phosphatase inhibition was less in tolerant compared to sensitive clones of A. odoratum and D. cespitosa
when grown in excess Zn (Cox and Hutchinson, 1980). Gibberellic acid added to the nutrient solution has been
shown to decrease the toxic effects of Zn in O. sativa seedlings (Nag et al., 1984). Gibberellic acid has an
enhancement effect on enzyme production, thus suggesting a pathway under which plant metabolism could
evolve to become to become more metal tolerant. Research with two cultivars of T. aestivum has shown that
total, apoplastic and symplastic Al are similar regardless of tolerance and solution Al concentration (Tice et al.,
1992). This suggests the requirement for a metabolically mediated tolerance mechanism, although it could easily
be one of the more standard mechanisms such as phytochelatins, organic acids or metallothioneins.
4.6.3 Conclusions on metal tolerance
It appears from the above information that the issue of metal tolerance in plants is far from resolved. Different
plant parts, species and metals appear to elicit different responses and possibly more than one response.
However there are certain mechanisms which appear to hold promise as being more widespread than others.
There is a body of evidence suggesting roots of many plants tolerant to Al exude organic acids. Exclusion of
metals via cell wall CEC has been sugested although there is contradictory evidence as to whether a high or low
CEC is desirable for tolerance. Plants tolerant to Cd, and possibly Zn, appear to utilise phytochelatins although
this may only be a transport system to sequester metals away in vacuoles. High cellular concentrations of
organic acids may have a role in metal tolerance, especially as the complexing agent in vacuoles. The role of
vacuoles in metal tolerance appears to be quite widespread for Zn, and also Cd. However, little testing has
occurred apart from these metals. Little evidence exists for metallothioneins or alterations of cellular
metabolism. However, this may just as likely be because of a lack of well designed and conclusive experiments
rather than these mechanisms not playing a role. Overall, not enough is known about metal tolerance to make a
pronouncement on the mechanism that any species uses without individual testing.
32
5. Methodology
5.1 Solution culture
5.1.1 Traditional non-renewed solution culture and dilute, renewed solution culture
There are two basic problems with traditional non-renewed solution culture methods. Firstly, to maintain high
enough concentrations of nutrients throughout the experiment, the initial solution concentrations are normally
much higher than typical soil solution concentrations (Asher and Edwards, 1978). Secondly, as the plants grow
and use nutrients, the concentration of nutrients in solution decreases. The combination of these factors means
that plants may experience basal nutrient concentrations ranging from toxic to deficient throughout the course of
an experiment (Asher and Edwards, 1978). In addition, across treatments, plants may experience vastly different
nutrient concentrations as growth differentials due to treatment effects result in differing nutrient uptake.
Non-renewed solution culture systems pose another set of potential problems in heavy metal research. The high
basal nutrient concentrations required may result in precipitation of the metals in high metal treatments. This is a
particular issue for metal-phosphates (Asher and Blamey, 1987). High basal nutrient concentrations may also
result in antagonistic or synergistic interactions in toxic effects of the metal being studied thus confounding
outcomes, for example Zn-Ca (Wallace and Abou-Zamzam, 1989) and Cu-P (Wallace and Cha, 1989)
interactions.
To minimise these problems a method known as Programmed Nutrient Addition was devised and has been used
successfully (Asher and Edwards, 1983). This method involves dividing the nutrient supply for the entire
experiment into a number of small doses given throughout the experiment. The method utilises already available
growth curves to estimate nutrient usage and thus the frequency and size of doses. Another similar method has
been devised by Ingestad (e.g.Ingestad and Agren, 1988) and involves growing plants with logarithmically
increasing inputs of materials intended to match the expected increase in plant demand.
5.1.2 pH
When studying the toxicity of metals in solution culture there are two important aspects to pH. Firstly, in trying
to model the soil solution it is important that the pHs utilised are representative of soils with toxic levels of
metals. Therefore, in general most metal toxicity studies should be undertaken at acid pHs (Gupta, 1972; Miles
and Parker, 1979; Lexmond, 1980; Suresh et al., 1987; Davis-Carter and Shuman, 1993; Watmough and
Dickinson, 1995). Secondly, the solubility and speciation of metal ions in solution is pH dependent. The exact
effect of pH will depend on the nutrient solution being used and metal being studied. In general, as the pH
increases the concentration of free metal-ion which can remain in solution decreases, and precipitates may form.
Precipitates are not available to plants growing in solutions. It should be remembered that precipitates are not
necessarily visible with the naked eye, and so authors such as Tam (1995) cannot use this method to claim a lack
of precipitates. Solution speciation programs such as GEOCHEM (Sposito and Mattigod, 1980) should be run
on any solution to be used to insure that metals will not precipitate, and hence, remain in available forms. From
Table 5.1 it can be seen that a variety of authors have used nutrient solutions which from which GEOCHEM
(Sposito and Mattigod, 1980) predicts a large proportions of the supplied metal would precipitate.
Table 5.1 Examples of nutrient solution metal concentrations, P concentrations and pHs employed by
various authors and the effective metal solution concentrations as determined by GEOCHEM (Sposito
and Mattigod, 1980)
Phosphorus
(mM)
pH
500 Cu
1.0
5.5
80 Cu
0.2
5.5
160 Cu
0.2
5.5
Metal (µM)
Metal
precipitated
(%) and form
83.9% as Cuphosphates
Soluble metal
(µM)
Reference
80.5
42.8% as Cuphosphates
78.1% as Cuphosphates
45.8
(GolanGoldhirsh
et
al., 1995)
(Ouzounidou et
al., 1995)
(Ouzounidou,
1994)
33
35.0
25 000 Cu
1.1
8
725 Cu
1
5.5
50 Zn
0.13
6.3
230 Zn
0.1
6.2
300 Mn
3
6.5
100.0% as Cuhydroxides
94.8% as Cuphosphates
33.5% as Znphosphates
38.5% as Znphosphates
35.2% as Mnphosphates
0.0
37.7
33.3
141.5
194.4
(Davis et al.,
1993)
(Schat
and
Kalif, 1992)
(Smirnoff and
Stewart, 1987)
(Berry
and
Wallace, 1989)
(Wang et al.,
1994)
It should be noted that other effects of high pH solutions can be precipitation of Fe-phosphates (Pettersson, 1976;
Wu, 1994; Golan-Goldhirsh et al., 1995), Fe-hydroxides (Russelle and McGraw, 1986; Smirnoff and Stewart,
1987; Galli et al., 1996) and Ca-phosphates (Davis et al., 1993; Wang et al., 1994). Iron and P deficiency can
both be induced by metal toxicity (Kabata-Pendias and Pendias, 1992; O'Sullivan et al., 1997) and Ca has a
restorative effect on metal toxicity (Kinraide, 1998). Hence, it is important to maintain adequate and equivalent
concentrations of these nutrients across metal treatments. Thus, it is important that these ions do not precipitate,
especially differentially across treatments. Methods of stopping precipitation are to keep the solution pH low,
supply the minimum amount of nutrient required to avoid a deficiency in control treatments, to maintain as low a
possible phosphate level in solution and check the speciation of solutions using software programs such as
GEOCHEM (Sposito and Mattigod, 1980). Nutrients can be kept low in solution by adding them sequentially
throughout the experiment, in methods like Programmed Nutrient Addition (Asher and Edwards, 1983).
5.1.3 Ionic strength
Numerous studies have shown that as the ionic strength increases the uptake of metal and therefore the apparent
toxicity of the metal in solution decreases (Vlamis and Williams, 1962; Wang et al., 1994; Parker et al., 1998).
The ionic strength effect has been attributed to competition between the metal and other ions for uptake. The
effect of high ionic strength may be due to the higher concentration of particular ions rather than the ionic
strength per se. Examples of particular ions which have been found to decrease metal uptake as the
concentration in solution increases include Ca, Mg and Cu on Zn uptake (Rashid et al., 1976), Ca, Mg, NH4, Fe,
sulphate and nitrate on Mn uptake (Vlamis and Williams, 1962; Elamin and Wilcox, 1986a; Elamin and
Wilcox, 1986b), and Ca and Fe on Cu uptake (Bowen, 1969). Research has also shown that the effect of
increased ionic strength could be attributed to the ameliorative effects of ions such as Ca (Wallace and AbouZamzam, 1989) or to the reduction of induced deficiencies (Wang et al., 1994), rather than direct competitive
effects on uptake. As well, high concentrations of some ions actually increase uptake of metals, such as Na and
phosphate on Mn uptake (Vlamis and Williams, 1962). While there is no one “correct” ionic strength, nutrient
solutions without confounding deficiencies/ toxicities of other ions and having comparable ionic strengths to
relevant soil solutions are likely to provide more meaningful results (see Table 5.2).
Table 5.2 Examples of soil solution ionic strengths
Soil
Seventeen soils
Ten soils
48 soils
Six weathered
soils
.
Origin
Western
Australia
South Pacific
Islands
Queensland
Queensland
Ionic strength (mM)
2.2-16.8 (topsoil)
1.8-12.1 (subsoil)
0.18-33.9
1.2-22.6 (topsoil)
<0.1-13.1 (subsoil)
1.47-12.31 (0-10cm)
Extractions method
Centrifuge
Reference
(Dellar
and
Lambert,
1992)
(Naidu et al., 1991)
Saturation paste,
vacuum filtered
Centrifuge
(Bruce et al., 1989)
Centrifuge
(Gillman and Bell, 1978)
5.1.4 Phosphorus
The amount of P in nutrient solutions used for the study of metal toxicity is important for two main reasons.
Firstly, to maximise the congruence with soil systems, P in nutrient solutions should be similar to that of relevant
soil solutions. Table 5.3 shows a range of soil solution P concentrations for world soils. Regular additions of P
as needed by the plants can be used to provide an adequate supply of P whilst subjecting the roots to the
generally low P concentrations found in soil solutions.
34
Table 5.3 Soil solution phosphorus concentrations for a cross-section of world soils
Soil
Origin
Seven basaltic soils
Six weathered soils
Four surface soils
Ireland
Queensland
?1
Two fertilised soils
?1
Four unfertilised
surface soils
Five woodland soils
Two grassland soils
Ten soils
?1
Britain
Soil solution P
(µM)
3.7-8.9
<20
0.6-1.3
0.6-1.0
0.6-1.0
1.9, 2.3
1.9, 2.3
5-55
6-58
33, 34
18-330
Extractions method
Reference
Displacement
Centrifuge
Column displacement
Centrifuge
Centrifuge, CCl4
Centrifuge, CCl4
Centifuge, EBA
Centrifuge, CCl4
(Benians et al., 1977)
(Gillman and Bell, 1978)
(Adams et al., 1980)
Centrifuge, Arklone P
(Campbell et al., 1989)
(Elkhatib et al., 1986)
(Elkhatib et al., 1987)
South
Saturation paste,
(Naidu et al., 1991)
Pacific
vacuum filtered
Islands
Seventeen soils
Western
<1-41
Centrifuge
(Dellar
and
Lambert,
Australia
1992)
1
Origin of soils not stated, however addresses of authors suggested the USA as place of origin
Secondly, at high metal concentrations in solution, P has a tendency to form insoluble metal-phosphates. Thus
rendering both the P and the metal being studied unavailable to plants. This has the potential to cause P
deficiencies in high metal treatments thus confounding results and also to result in added metal being a
misrepresentation of metal experienced by the plant (see Table 5.1). Solution speciation programs such as
GEOCHEM (Sposito and Mattigod, 1980) can be used to determine if metal-phosphate precipitates will form in
a given solution.
5.1.5 Iron and chelates
Iron readily forms insoluble Fe-oxides and Fe-phosphates in solution, with these being unavailable to plants
(Halvorsan and Lindsay, 1972; Cline et al., 1982). To keep Fe available and prevent deficiency, Fe is often
added to nutrient solutions in chelated form. However, other metals have high or higher affinities for chelates
and may displace Fe from the chelate complex (Gunn and Joham, 1963; Halvorsan and Lindsay, 1972). Many
researchers have shown that chelates reduce the plant uptake of metals in nutrient solutions (DeKock and
Mitchell, 1957; Wallace and Mueller, 1962; Halvorsan and Lindsay, 1977; Bachman and Miller, 1995).
Hence, adding chelated Fe to nutrient solution may result in misrepresentations of the availability of the metal
ions being studied. From these studies came the idea that while chelates can facilitate the movement of the metal
to the plant it is the free-ion in solution which is available to plants and should be measured (Halvorsan and
Lindsay, 1977). However, Cu and Zn uptake at low to medium solution concentrations has been shown to be
increased by the presence of EDTA (Checkai et al., 1987). Work by Srivastava and Appenroth (1995) has
shown that the uptake of metals in chelated solutions is slightly higher than the free-ion concentration,
suggesting that chelates are a confounding factor in nutrient absorption, that the chelate metal-solution metal
system stays in equilibrium with solution metals being replenished as plants and/or precipitation removes them
from solution. Hence, while the free-ion activity is a good indicator of plant availability, it is still only an
estimate and chelates are a confounding factor in metal toxicity studies. Therefore, where metal concentrations
are being studied it is important that minimal amounts of chelates are added to solution.
This then poses the question of how to supply adequate Fe to plants. Metal-chelate complexes can be minimised
by making sure that minimal amounts of Fe-chelates are added and that the solution pH remains low (Wallace,
1962c; Wallace, 1962a; Gunn and Joham, 1963). However, the benefit of low pH varies for different chelates.
For example, EGTA has been shown to complex virtually all Cu added at pHs from 4 to 9, wheras, for EDTA it
is only above pH 5.5 that most of the Cu is complexed (Halvorsan and Lindsay, 1972).
Iron deficiencies can be successfully prevented by painting or spraying plant leaves with Fe solutions (Halvorsan
and Lindsay, 1977; Modaihsh, 1997). Care must be taken to get an even supply of Fe over the plant and to
reapply as new leaves appear as Fe is not readily translocated within the plant (Wallace, 1962b). However,
metal toxicities frequently result in induced Fe deficiency in plants. (Halvorsan and Lindsay, 1977; Wallace and
35
Abou-Zamzam, 1989; Ren et al., 1993; O'Sullivan et al., 1997). Therefore, painting of plant leaves with Fe
solutions could lead to misrepresentations of the level of toxicity being experienced by a plant.
The Fe-chelate issue can not be overcome by adding smaller amounts of Fe-chelate throughout the experiment as
the chelate may accumulate in the solution throughout the course of the experiment. Metal speciation could also
be estimated or measured and the level of free metal–ion used as the treatment not the total metal concentration.
However, by the very nature that the authors in Table 5.4 were able to induce Cu toxicity, especially Wallace
(1989), suggests that while chelated Cu may not be as toxic as the free ion form it is definitely toxic to some
degree. This toxicity may be as a result of metal being removed from the chelate near the plant root.
Table 5.4 Examples of nutrient solution metal and chelate concentrations employed by various authors
and the soluble metal solution concentrations as determined by GEOCHEM (Sposito and Mattigod, 1980)
Total
Metal
(µM)
10 Cu
25 Cu
0.2 Cu
50 Cu
800 Cu
Chelate
EDTA
EDDHA
Chelate
concentration
(µM)
10
50
pH
% metal as
M-chelate
5.5
8
4
5.0
5.0
5.0
99.4
100.0
98.6
100.0
35.8
97.7 (2.2)
17.9
17.9
800 (17.9)
10 Cu
HEDTA
HEDTA
EDTA
(HEDTA)
citrate
50
6
98.9
80 Zn
40 Zn
EDTA
EDTA
20
20
100
6
6
24.2
46.2
97.3
Soluble metal
(µM)
0.06
0.00
0.04
0.0
32.1
0.8
0.11
60.6
18.5
1.1
Reference
(Adalsteinsson, 1994)
(Wallace, 1989)
(Taylor and Foy, 1985)
(Pettersson, 1976)
(Fontes and Cox, 1995)
(Fontes
and
Cox,
1998b)
The chelate issue should be thoroughly thought out if research on the effects of chelates is to be undertaken. For
example, from the work of Taylor and Foy (1985) (see Table 5.4) it can be demonstrated that their ‘nonchelated’ treatments (0.2 and 50µM as examples) were in fact confounded by the HEDTA chelate added to
maintain Fe in solution.
36
6. Metal toxicity and Australian native plants
6.1 Information on the tolerances and responses of Australian plants to
excess Cu, Mn and Zn
Australian has many unique species of plant, and so, it is likely that Australian species would have different
tolerances and responses to excess metal than exotic species. Unfortunately, this subject has received little
attention.
In a study designed for the horticulture industry, the effect of lining pots with carbonates of Cu and Zn, as well
as Co, Pb, Ni and Sr, on root growth of Anigozanthos flavidus (kangaroo paw) was studied (Baker et al., 1995).
The aim was to find a successful method of inducing root pruning in container-grown plants, thus preventing pot
bound roots. Growth was found to be reduced by the use of metals however because of the design of the
experiment little can be said in a quantitative manner. In a further study on A. flavidus (Kaub et al., 1998), Cu,
Zn, and Co coated pots were found to induce necrosis, reduce root concentration of K, and decrease the mitotic
index in root-tips near the container wall, as well as increasing the root to shoot ratio.
The effects of metals in emissions from a coal-fired power plant on tissue concentration in E. crebra and
Eucalyptus moluccana (grey box) were analysed (Murray, 1984). Foliar concentrations of Mn and Zn ranged
from 199-976 and 8-20 mg.kg-1 in E. crebra, respectively, and from 44-1002 and 10.5-42.5 mg.kg-1 in E.
moluccana, respectively. The authors did not mention any toxicity symptoms, but, concentrations of Mn above
500 mg.kg-1 can be toxic to exotic species and tolerant species can withstand concentrations greater than 1000
mg.kg-1 (Kabata-Pendias and Pendias, 1992). As a result of the lack of information on the tolerances of
Australian plants to metal toxicity it is possible, but not conclusive, that these species were experiencing
deleterious effects from the excess metals.
A study compared Cu toxicity in Australian trees i.e. Banksia ericifolia (heath banksia), Casuarina distyla (sheoak) and Eucalyptus eximia (yellow bloodwood) (Mitchell et al., 1988). Concentrations of Cu in the soil found
to reduce growth by 50% were 205 mg.kg-1 for C. distyla, 560 mg.kg-1 for E. eximia and 610 mg.kg-1 for B.
ericifolia. However, only total soil metal concentrations were measured and so results cannot be extrapolated
outside the particular sandy loam soil used in the experiment. Stunted growth and chlorosis were noted as Cu
toxic symptoms and excess Cu was found to delay and reduce final percentage emergence.
Eucalyptus saligna (Sydney blue gum) has been shown to be more tolerant of high Mn soils than E. gummifera
(Winterhalder, 1963). Eucalyptus gummifera plants showed Mn toxicity symptoms of small chlorotic leaves of
distorted shape and regular death of the terminal bud at 2040 mg.kg-1 leaf Mn but not at 510 mg.kg-1. Spraying
chlorotic leaves with Fe solution resulted in recovery from the chlorosis, suggesting induced Fe deficiency. In
comparison, E. saligna plants showed no symptoms of Mn toxicity at leaf concentrations up to 4250 mg.kg-1. In
another study on Mn toxicity in M. integrifolia, leaf Mn concentration was often not toxic at concentrations in
the leaves of greater than 1000 mg.kg-1 (Hue et al., 1987).
The most comprehensive study of Zn toxicity on an Australian tree (A. auriculaeformis) found a critical nutrient
solution concentration of 0.7 mg.l-1 (9 µM) Zn. This corresponded to a Zn concentration of 143 mg.kg-1 in the
shoots.
It is important that we understand the toxicity responses of Australian trees to metals so that we can effectively
manage metal contaminated areas. Therefore, research is needed which quantifies results giving information on
tolerances across a range of metal contamination levels, and measuring bioavailable metal and tissue
concentrations in index leaves.
6.2 Within species genetic variability and the significance for metal
tolerance
Most research on metal toxicity has considered crop and pasture species (e.g. Brookes et al., 1981; Chino and
Baba, 1981; Cox, 1990; Parker et al., 1990; Fontes and Cox, 1995), which have been bred for minimal genetic
diversity (Raven et al., 1986). In comparison, Australian plants have evolved over millions of years in an
37
unpredictable and ever-changing environment where increased genetic diversity, hence, adaptability, has often
been advantageous (James, 1981; Coates and van Leeuwen, 1996).
Genetic diversity has two implications for metal tolerance in Australian plants. Firstly, for species which have
naturally grown over a wide geographic range in Australia it is recognised that environmental clines over the
species range may result in localised populations with variations in genetic make-up (Coates and van Leeuwen,
1996). These local populations are known as provenances (Coates and van Leeuwen, 1996; Matyas, 1996).
Inter-provenance variation in Australian plants has been documented for varying tolerances to many
environmental variables including waterlogging, salinity, alkalinity, frost and drought (Merwin et al., 1995;
Farrell et al., 1996; Li et al., 2000). Although no research has been undertaken in the area of metal toxicity, it is
feasible that some provenances of Australian species would have better tolerances to metals than others. Hence,
testing one provenance for metal tolerance may not incorporate the true species diversity and this should be
taken into account when looking at studies where only one provenance has been tested.
Secondly, considerable within provenance genetic variation exists in some Australian species (Merwin et al.,
1995; Farrell et al., 1996). This intra-population genetic variability could mean that while the average plant
within a provenance may not be tolerant of a given metal supply a considerable proportion might be. Hence,
average results should be taken as just that, “average”, and it is important to understand and utilise the natural
variation within a population to optimise rehabilitation of contaminated areas.
38
7. Final comments
Metal toxicity issues in plants and soils are a significant problem throughout the world, including Australia. It is
only by understanding the relationships between bioavailable metal fractions in the soil and plant responses to
metals that we can make decisions regarding metal toxicity in plants. At present, research has still not
conclusively defined the bioavailable metal fraction nor do we fully understand the processes behind metal
toxicity responses in plants. As well, many poorly designed experiments have been undertaken, where
confounding effects such as that of pH on bioavailability in soil, or precipitation of metals in solution culture,
has reduced the precision of and confidence in results obtained. Little research, particularly of a quantitative
nature, has been undertaken on Australian plants. If we are to understand the effects of excess metal on
Australian plants it is important that well designed, quantitative research be undertaken. This will enable the
rehabilitation of metal contaminated areas with appropriate Australian species, allow identification of metal
toxicities when they occur and allow for the effective regulation of metal emissions.
39
8. References
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geometry and nutrient distribution. Journal of Plant Nutrition. 17, 1501-1512.
Adams, F., Burmester, C., Hue, N.V. and Long, F.L. (1980). A comparison of column displacement and
centrifuge methods for obtaining soil solutions. Soil Science Society of America Journal. 44, 733-735.
Aery, N.C. and Jagetiya, B.L. (1997). Relative toxicity of cadmium, lead, and zinc on barley. Communications in
Soil Science and Plant Analysis. 28, 949-960.
Agbenin, J.O., De Abreu, C.A. and van Raij, B. (1999). Extraction of phytoavailable trace metals from tropical
soils by mixed ion exchange resin modified with inorganic and organic ligands. Science of the Total
Environment. 227, 187-196.
Aidid, S.B. and Okamoto, H. (1992). Effects of lead, cadmium and zinc on the electric membrane potential at the
xylem/symplast interface and cell elongation of Impatiens balsamina. Environmental and Experimental
Botany. 32, 439-448.
Alam, S., Kamei, S. and Kawai, S. (2000). Phytosiderophore release from manganese-induced iron deficiency in
barley. Journal of Plant Nutrition. 23, 1193-1207.
Alva, A.K., Blamey, F.P.C., Edwards, D.G. and Asher, C.J. (1986). An evaluation of aluminum indices to
predict aluminum toxicity to plants grown in nutrient solutions. Communications in Soil Science and Plant
Analysis. 17, 1271-1280.
Amberger, A. and Yousry, M. (1988). Study on the effects of increasing manganese concentrations in nutrient
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