Exploiting plant metabolism for the phytoremediation of persistent herbicides

Persistent Herbicides Phytoremediation

Phytoremediation: Persistent Herbicides

Exploiting Plant Metabolism for the Phytoremediation of Persistent Herbicides Julian O. D. Co leman 1, Car la Frova 2, Peter Schr6der 3" and Michel Tissut 4

1 School of Biological and Molecular Sciences, Oxford Brookes University, Gipsy Lane, Oxford OX3 0BP, UK 2Department of Genetics and Microbiology, University of Milano, Via Celoria 26, 1-20133 Milano, Italy 3 GSF-Institute for Soil Ecology, GSF Research Centre for Environment and Health, D-85764 Neuherberg, Germany 4Laboratoire Ecosyst~mes et Changements Environnementaux, Universit6 J. Fourier, BP 53, F-38041 Grenoble cedex 9, France

* Corresponding author ([email protected]

D O h h t t p : / / d x . d o i . o r g / 1 0 . 1 0 6 5 / e s D r 2 0 0 1 . 0 9 . 0 8 4 . 5

A b s t r a c t . Weed control by herbicides has helped us to create the green revolution and to provide food for at least the majority of human beings living today. However, some herbicides remain in the environment and pose an ecological problem. The present review describes the properties and fate of four representative herbicides known to be presistent in ecosystems. Meta- bolic networks are depiced and it is concluded that removal of these comopounds by the ecologically friendly technique of phytoremediation is possible. The largest problem is seen in the uptake of the compounds into suitable plants and the time needed for such an approach.

Keywords: Atrazine; chloroacetanilides, isoproturon; metolachlor; persistent herbicides; phytoremediation; plant metabolism

I n t r o d u c t i o n

Over the last 50 years weed control by herbicides has become an integral part of commercial agriculture and because of its cost-effectiveness, herbicide use is likely to remain an important agronomic strategy for the foreseeable future. In order to ensure that agrochemical compounds produce no 'unacceptable risk' to human health or the environment, all

herbicides are subject to an approval procedure under EU and national legislation that imposes detailed controls on the use of each product. However, one of the inevitable re- percussions of the widespread use of herbicides is contamination of soil, surface water and ground water by parent compounds or their degradation products. Heavy environmental contamination by herbicides may arise from indus- trial point sources such as accidental spillage during pro- duction, storage and transport, or as leachates from former dumping sites. In contrast, sources of pollution arising from agricultural uses of herbicides (e.g. crop spraying) are considered to be diffuse as the compounds are distributed over large areas. During herbicide spray application less than five percent of the compound will reach the intended target due to spray drift, volatilisation or washout. Herbicides that are sprayed on the foliage of crops and weeds (e.g. post-emergence herbicides) reach surface waters via the air and soil. Herbicides that are applied to the soil (pre-emergence herbicides) through leaching or run-off may reach deeper soil lay- ers or adjacent surface water and groundwater aquifers.

The concentration and persistence of herbicides in the environment depend on several factors including crop and soil type, application rates, degradation rates and the physico- chemical properties of the compound (Fig. 1). Good examples of the latter are the properties of soil applied pre-emer-

Fig. 1 : Differential entry paths of per- and post-emergence herbicides into agricultural ecosystems

18 ESPR - Environ Sci & Poliut Res 9 (1) 18 - 28 (2002) �9 ecomed publishers, D-86899 Landsberg, Germany and Ft. Worth/TX, USA �9 Tokyo, Japan �9 Mumbai, India �9 Seoul, Korea

Phytoremediation Persistent Herbicides

gence herbicides. Pre-emergence herbicides are designed to kill emerging weed seedlings soon after germination. As weed germination can occur over extended periods pre-emergence herbicides are formulated to bind to upper soil components and to degrade slowly. These design properties inevitably contribute to the persistence of these compounds.

Herbicides differ in many significant ways from other organic pollutants and these differences have important impli- cations for their removal or detoxification during phytoremediation. Herbicides by their very nature are designed to kill unwanted plants and do so by blocking vital biochemical and physiological processes. Because the processes affected by herbicides may be common to a wide range of plant species and to other non-target organisms, herbicides can be severe pollutants even at low concentrations. Typically herbicides are designed to be selective. This means that some tolerant crop plants can metabolise particular herbicides or a family of herbicides. In addition some weed species have acquired herbicide resistance by point mutations at the target site or by the selection of genes that lead to enhanced metabolism of the herbicide. Consequently for phytoremediation, plants capable of removing and trans- forming herbicides are already known. However, a formulated herbicide consists of a mixture of biologically active ingredients and additive chemicals such as solvents and surfactants. These additives are devised to enhance plant uptake and may improve the chances for phytoremediation. On the other hand, some active compounds may be present as a racemic mixture of isomers and the enantomeric selectivity of plants may affect the process of phytoremediation by removal of only one isomer.

In many countries all herbicides are subject to approval and regulatory procedures and standards for use are set interna- tionally (e.g. EU legislation and W H O recommendations) and nationally. The European Commission's Directive on Drinking Water 80-778-EEC (1980) established a maximum allowable concentration of 0.1 ~tg/L for individual pesticides and 0.5 ~tg/L for total pesticides. Many European countries have adopted these statutory limits.

In Europe (EU countries), 33 herbicides being approved for use have been detected and monitored in ground water. Of these, approximately 12 compounds from 4 chemical families of herbicides have been detected at some sampling sites with concentrations >0.1 ~tg/L. and were classified as important pollutants. Families of herbicides and examples of member compounds are shown in Table 1.

In this review we describe the general properties of four of these compounds (atrazine, metolachlor, isoproturon and

Table 1 : Families of herbicides and examples of member compounds

2,4-D) and their uptake and metabolism in plants. On the basis of these case histories we assess the prospects of using phytoremediation technology for the in situ treatment of herbicide-contaminated water, sediments and soil.

1 Atrazine

The triazines are a widely used family of herbicides that were developed during the 1950s and 1960s. The characteristic structural feature of triazine herbicides is the symmetrical 1,3,5- triazine ring (Fig. 2). These compounds are photosynthesis inhibitors, acting on the electron transfer between photosystem II and the plastoquinone pool, in chloroplast thylakoids (Good 1963). Since they were first introduced, several triazine derivatives have been developed with selectivity for use in many important crops such as maize, sorghum, sugar cane, vines, oil palms, coffee and in forestry. The wide spread use of triazines is exemplified by the case of atrazine whose application reached 100,000 tons per year by the end of the eighties. However, due to several resistance problemsand connected to its high persistence, atrazine was banned in some european countries at the beginning of the nineties.

Atrazine (Fig. 2, Table 2) is a selective systemic herbicide used for the pre-and post-emergence control of annual grasses and broad-leaf weeds in a variety of cultivated crops. However the major use of atrazine is the pre-emergence control of weeds in maize and sorghum. Atrazine absorbed by foliage and roots, is translocated acropetally in the xylem and accumulates in the apical meristems where it develops its herbicidal action.

~N I~CI (CH3~CHNH~T~NyC' N~N NHCH2CH3

A. B. Fig. 2: Chemical structures of triazines. A. 1.3.5-triazine nucleus. B. Atrazine

Table 2: Physical properties of Atrazine

Molecular formula CsH14CINs

Molecular weight 215.7

Water solubility 33 (22 C; pH 7)

Vapour pressure 0.039 mPa (25 C)

Log Kow 2.5 (25 C) (non-ionised form

PK, 1.7 (base)

Family Common Name

Triazines Atrazine

Chloroacetanilides I Metolachlor

Ureas I Isoproturon

Aryloxyalkanoic acids I 2,4-D

Compound I Chemical name (IUPAC)

I 6-chioro-N2-ethyl-N4-isopropyi- 1,3,5-t riazine-2,4-diamine

I -chloro-N-(2-ethyl-6-methylphenyl)-N-(2-methoxy-1-methylethyl)acetamide

I -(4-isopropylphenyl)- 1 !1 -dimethylurea

(2,4-dichlorophenoxy)aceticacid

ESPR - Environ Sci & Pollut Res 9 (1) 2002 19

Pers is ten t H e r b i c i d e s Phy to remed ia t ion

1.1 Atrazine metabolism in plants

In corn, atrazine is metabolised by three major pathways:

a) dechlorination of the C 2 followed by hydroxylation takes place in the presence of a secondary metabolite (Lamoureux et al. 1998) that is accumulated to high concentrations in maize. This chemical mechanism was first described by Castelfranco et al. (1961) and its role demonstrated by Raveton et al. (1997). This process is a very active metabolic reaction in the young seedling. The hydroxylated atrazine is practically devoid of herbicidal activity.

b) In the aerial parts of the maize plant, a large pool of glutathione-s-transferases catalyses the formation of glutathione conjugates with several xenobiotic substrates (Marrs 1996). One isoenzyme of this pool catalyses glutathione conjugation to the C a of atrazine to give a highly hydrophilic derivative (Shimabukuro et al. 1971).

c) In many species N-dealkylation occurs catalysed by an N- dealkylase (Lamoureaux et al. 1997), leading to the formation of the hydrophilic derivatives N-deethylatrazine and N-deisopropylatrazine. These derivatives are not phy- totoxic. In maize leaves the N-dealkylase seems to be much less active than the glutathione-S-transferase.

1991) sometimes at levels in excess of the regulatory limit (Bintein and Devillers 1996). Moreover, atrazine traces can be detected in the atmosphere and in rain (Glotfelty 1978, Glotfelty et al. 1990, Goolsby et al. 1995, Bester et al. 1995). Although atrazine is neither a stable compound in the biosphere nor an accumulated one, the conditions of its world- wide agronomic use has led to its appearance in several environmental compartments. Therefore for all practical pur- poses it may be considered as persistent pollutant.

Among the most interesting possible uses of phytoremediation of atrazine contamination would be: 1) To prevent leaching to groundwater from unplanted

fields, for instance from the end of August to winter in maize cultivation. This would also help to attenuate ni- trate leaching;

2) to prevent surface run-off to the rivers either by main- taining buffer zones planted with species capable of me- tabolising atrazine or by cultivating detoxifying aquatic plants in the ditches, storm basins or wetlands.

Of course, the choice of plants for this purpose must be some- what restricted to monocotyledoneous species, as most broadleaf plants are killed by the compound.

1.2 Atrazine degradation in soils

Several studies carried out on seedlings under controlled conditions or in the field show that the metabolic potential of maize for atrazine is less than 100 g ha -1 year -1 for a field with 90,000 maize seeds ha -1 and treated with 1500 g atrazine ha -1 (Tasli et al. 1996). Under agronomic conditions, most of the metabolism of atrazine occurs in the upper layer of the soil and is carried out by micro-organisms (Mandelbaum et al. 1995, Mirgain 1993, Shao and Behki 1995, Anderson and Coats 1995, Ostrofsky et al. 1997, Masaphy et al. 1996). Generally, the first step in soil microbial metabolism is N-dealkylation leading to the formation and accumulation of the metabolites N-deethylatrazine, N- deisopropylatrazine and N-didealkylatrazine in the soil. In maize fields treated repeatedly with atrazine, the capacity of the soil to metabolise this herbicide increases as the result of selective pressure on the microflora. It could be anticipated that all these transformation processes might ensure the rapid disappearance of atrazine from the treated areas. However, more than thirty years of repeti- tive usage of this herbicide has led to contaminating traces of this product and its metabolites in water (Buffaut 1992) (Table 1, p. 17), food (Gunderson 1995, Wittmann and Hock

2 Isoproturon

Most members of the urea group of herbicides are compounds with dimethyl or methyl/methoxy substituent on the terminal N atom (Fig. 3, Table 3).

The phenylurea herbicide, isoproturon (IPU), [N-(4-isopropylphenyl)N',N'-dimethyl urea], is mainly used as post emergence herbicide in the control of grasses (e.g. blackgrass) and dicot weeds in cereals at application rates of up to 2 kg/ha. The mode of action is the plastoquinone antagonism at photosystem II. Its selectivity is based upon enhanced rhe- tabolism in crops compared to the susceptible weeds.

o ,R 3 R ~ N H - C - - N

�9 ~2 Y R4 . ,~--~ o ,CH~

P 0 ~ N H - - C - N , c H 3

A. a .

Fig. 3: Chemical structures ureas. A. Urea nucleus. B. Isoproturon

Table 3- Some chemical properties of the herbicide, isoproturon (from Perkow, 1988, Behrend and Br0ggemann 1993, modified)

Formula C12HlaN20

Molar mass 206.29 g/tool

Melting point/boiling point 151-153~ distillable

Stability Stable against light and acids, hydrolysis in the presence of strong bases

Vapour pressure 0.31 10 -'~ mbar at 77~ 0.17 10 -2 mbar at 90~

Solubility 170 mg/L in water at 25~ good in all solvents

Partition coefficients: aWwater; log kow

Sorption to soil (equilibrium coeff. K d (cm3/g)

KAw = 3.9 104, Iog Kow = 2.3

sand: 0.31; loam: 0.92

20 ESPR - Environ Sci & Pollut Res 9 (1) 2002


In animals, IPU exhibits mutagen and potential hepatocarcino- gen properties. Connected to its ability to reach groundwater and surface waters easily (Schiilein et al. 1997), it is a compound with possible negative impact on human health.

Typically, IPU is applied to the soil as pre-emergence herbicide. Phenylureas are taken up into the plant via the leaves or the roots, but root uptake is the predominant uptake path. After transport in the xylem, the chemicals accumulate in the edges of the leaves where first necroses are observed. In the soil, IPU seems to persist for longer periods of time. In several investigations, autumn applications of IPU were still found after 10 weeks (30%; Kulshrestha 1982), 30 weeks (2-8%; Moss 1979), and a similar period (40-80%; Anon 1979). However, heavy rainfall may render the compound and its metabolites mobile as has been shown in a report examining the migration of IPU on a bavarian research farm. It was found that IPU was translocated to depths down to 170 cm at concentrations exceeding 4 lag/L and its metabolites at concentrations of 1 lag/L.

In any case, the breakdown of IPU in soil follows first order kinetics wi th respect to the herbicide concent ra t ion (Kulshrestha 1982). It is significant from these reports, that the presence of plants (wheat) on the sites does not contribute to an increased breakdown of the compound. This increased persistence is apparently due to an increase in the strength of adsorption to soil matrices with time (Blair et al. 1990). The adsorption seems to be a function of organic matter but not governed by the clay content of the soils in- vestigated (Yadak and Malik 1988). Similarly, isoproturon uptake into wheat plants was found to be more rapid from soils with low sorption coefficients ( K a ) than from those with high K a (Blair et at. 1991).

From these reports it seems as if the desorption from the soil and the subsequent uptake of the parent compound into the plant represented a major barrier to remediation. Ex- periments on the addition of Triton X-100 to the applied herbicide indicate that increased solubility leads also to higher penetration into tissues (Ponte-Freitas et al. 1991).

Once inside the plant, IPU tolerance is conferred by metabolism rates. Generally N-demethylation or C-hydroxylation are typical plant reactions on phenylurea herbicides. This has been shown for monuron and diuron (Frear et al. 1972), and chlortoluron (Gross et al. 1979). Accordingly, also plant metabolism of IPU is mediated by P450 monooxygenases. As first steps of metabolism hydroxylation and demethylation have been shown (Haas 1997). Although both reactions are catalyzed by monooxygenases, evidence has been presented that two distinct P450s are responsible for these steps. It is significant to note that the product of the demethylation reaction, Monodesmethyl-isoproturon, exhibits residual phytotoxicity, whereas the metabolite, 2-OH-IPU, is non toxic. In ki- netic studies with cell cultures it could be shown that wheat and soybean differ exactly in this respect in their detoxification capacity (Gliif~gen et al. 1999). Whereas soybean cell cultures accumulate the toxic desmethyl IPU, predominantly hydroxylated metabolites are found in wheat. In a later step,

wheat builds up 2-OH-monodesmethyl-IPU as intermediary metabolite. In both cultures, 45 to 54% of the applied IPU were transformed to more polar metabolites within 48 hrs. After 7 days, wheat metabolized up to 87% of the parent compound. Contrary to the wheat cell culture studies, mi- crosome preparations from field grown wheat plants reveiled the desmethyl-IPU as major metabolite. This contradiction is not well understood and ha's not been unravelled by Gl~it~gen and coworkers (1999).

Whereas the closely related phenylurea chlortoluron is further me tabo l i zed to glucosides , no evidence for a glycosidation of IPU or its metabolites has been detected in wheat and soybean (Glfit~gen et al. 1999). It may be dis- cussed that glucosyl transfer is sterically inhibited in the case of isoproturon. Consequently, the amount of bound residues formed in the plant is also very small (below 5%).

As for phytoremediation, it has to be concluded that the removal of IPU from contaminated land seems to be governed mainly by its bioavailability in various soil types and by the uptake into the plants. Once metabolised by soil bacteria, the uptake of break down products of IPU has to be regarded as to be poor. However, special attention has to be focused on the fact that phenylureas are metabolized to chloroanilins, which might themselves be a threat to biosphere. In water, there might be a net uptake of the isoproturon observed depending on the conditions set. Fuertet-Mazel and co-workers (2000) found that IPU was bioconcentrated in the freshwater macrophytes Elodea densa and Ludwigia natans up to 60fold. However, this process might be exploited for bioremediation only for a short period of time as the plants showed massive growth reduction.

In conclusion, phenylureas tend to become persistent in soils with high organic carbon contents. Even if they are biode- gradable, a good proportion may remain present in the soil in the form of chloroanilins which have their own toxicity and persistence problems (see review of Harvey et al. this issue). For phytoremediation, most cereals would be good targets as many of them are tolerant due to their metabolic features. However, uptake of the isoproturon into the plant root has to be facilitated. Hence experimental and practical approaches for phytoremediation could include the use of biotensides such as rapeseed oils and esters. Another possibility that has not been experimentally proven so far is to set the area under con- sideration aside and leave it as fallow, but cut and remove biomass. Under these conditions, organic matter will dimin- ish with time and phenylurea herbicide residues could become accessible to soil organisms or plant roots.

3 Metolachlor

The chloroacetanilides are a large group of pre- or early post- emergence herbicides, active against a wide spectrum of annual grasses and broad-leaved weeds. Metolachlor is one of the most extensively used chloroacetanilides for selective weed control in maize, sorghum, soybeans, peanuts, sugarcane, cot- ton, potatoes, vegetables and woody ornamentals.



N-~,3-CH2--CI

R2

Et O ~-CHz--Cl

N" / - - ( ~ ' - "CH--CH;t-OMe

Me Me

A. g. Fig. 4: Chemical structure of chloroacetanilides. A. Chloroacetanilide nucleus. B. Metolachlor

Fig. 4 shows the chemical structure of cloroacetanilides in general (Fig 4a) and of metolachlor in particular (Fig 4b). The presence of the cx-chlorine atom is necessary for herbicide activity (Le Baron 1988).

Chloroacetanilides are active as general plant growth inhibitors (Fedke 1982). Inhibition of protein and lipid synthesis has been proposed as a mechanism of phytotoxicity for all chloroacetanilides, but their primary biochemical mode of action is still not defined. Moreover, the physiological effects on plants depend on the herbicide used, plant species, tissue type, herbicide concentration and treatment duration (reviewed in Le Baron 1988). Metolachtor has been shown to affect seed germination, root and shoot development, cell division, protein and lipid synthesis, membrane permeabil- ity (Fuerst 1987, Le Baron 1988, Matthes et al. 1998, Schmal- fuss et al. 2000).

The differences in tolerance to chloroacetanilides observed between plant species may be due to differences in either uptake/translocation of the herbicide or in its metabolism. In general, however, the major mechanism responsible of differential tolerance to chloroacetanilides is thought to be differential metabolism or rate of metabolism. The major route for chloroacetanilides metabolism starts with a conjugation reaction with glutathione (see paragraph on metabolism). Accordingly, tolerance has been correlated with a higher availability of reduced glutathione and/or with the expression of specific glutathione S-transferase isozymes with a greater activity towards defined herbicidal molecules (Fuerst 1987, O'Connell et al. 1988, Farago et al. 1993, Cottingham and Hatzios 1992).

Several major crops, including corn and soybean, are generally tolerant to metolachlor doses effective in weed control. However, intraspecific variability for tolerance is common: in corn, for instance, significant differences in chloroacetanilides tolerance both between inbreds and hybrid varieties have been detected (Sari Gorla et al. 1993, Cottingham et al. 1993, Cottingham & Hatzios 1992), meaning that also in generally tolerant species these herbicides are not always completely selective. To overcome this problem safeners can be used in conjunction with the herbicide itself, in order to enhance the tolerance of the crop. Safeners like cytometrinil, flurazole, oxabetrinil, fluxotenim, benoxacor have been suc- cessfully used with metolachlor in sorghum and other non completely tolerant crops. The biochemical and molecular mechanisms of action of herbicide safeners are not fully understood yet. However, at least for chloroacetanilide

Table 4: Physical properties of Metolachlor

Molecular formula C~sH=CINO=


Water solubility 488 mg L -~ (25 C)

Vapour pressure 4.2 x 10 -~ Pa (25 C)

Log Kow 2.9 (25 C)

safeners, it seems clear that they mainly act by enhancing metabolic detoxification (Hatzios 1997, Jepson et al. 1994, Irzyk and Fuerst 1997, Wu et al. 1996, Han and Hatzios 1991) (Table 4).

3.1 Metolachlor uptake and translocation

Plants absorb chloroacetanilides through both roots and shoots. Corn absorbs metolachlor mostly through shoots, and variability in uptake between genotypes has been shown (Kreuz et al. 1989). Translocation of metolachlor and alachlor in plants is mostly by acropetal movement, with some limited basipetal movement (Obrigawitch et al. 1980). [14C]metolaclor applied to roots of corn seedlings grown in nutrient solution was taken up rapidly. After one week 30% of the radioactivity was found in aerial parts and 70% still in roots (Gross, reported in Le Baron 1988). Corn grown in a field plot treated preemergence with [14C]metolachlor showed, at maturity, some radioactivity in the stalk and al- most none in seeds (Sumner & Cassidy, reported in Le Baron 1988). When metolachlor was applied to seedling coleoptiles, most of it (65%) remained in the tissue, some (24%) was translocated to mesocotyls, and only marginal proportions could be detected in leaves, roots and seeds (Kreuz et al. 1989, Cottingham and Hatzios 1992).

3.2 Metolachlor metabolism

Detoxification of chloroacetanilides in planta has been shown to occur by different metabolic pathways. By far the major is the glutathione S-transferase (GST) mediated conjugation of the herbicide to reduced glutathione, GSH, (step 1), followed by vacuolar compartmentation of the conjugate through the action of a GS- ATP-dependent pump (step 2). Evidence exists for further metabolization of the glutathione conjugate in the vacuole (step 3) and possibly final excretion of the metaboIites with inclusion in the cell wall (step 4). The first two steps are so far the best characterized and recognized as of major importance. GSH, GSTs and the GS-pump are therefore critical for chloroacetanilides detoxification.

GSTs are a very abundant and ubiquitous class of enzymes encoded by a large multigene family. They are active as homo- or hetero-dimers and catalyse the conjugation of electrophilic and frequently hydrophobic toxic compounds with glutath- lone to form less or non toxic peptide derivatives. As for chloroacetanilides, they operate the glutathione (or homo- glutathione in some legumes) displacement of the electrophilic chlorine atom, thus drastically reducing their herbicidal activity.



Plant GSTs classification on a functional basis has proved difficult and confusing, mainly because several members of the family are multifunctional enzymes that show broad substrate specificity. A recent classification based on aminoacid sequence similarity and intron-exon organization of the encoding genes (Dixon et al. 1998, Edwards et al. 2000) has recognized four classes of plant GSTs: Phi (F), Zeta (Z), Tau (U) and Theta (T). The Phi class is plant specific and includes most of the herbicide-detoxifying GSTs; their genes contain 3 exons. The Tau class, 2 exons, is also plant specific and includes many auxin-induced GSTs. These two classes account for the majority of plant GSTs. The other two classes, Zeta (10 exons) and Theta, are close to mammalian Zeta and Theta classes respectively, but are much less abundant in plants. GSTs with similar function and substrate specificity often belong to the same class, but sub- stantial overlapping has been observed. For instance, in maize, alachlor and metolachlor are mainly detoxified by four phi GSTs, ZmGSTFI-1, F1-2, F2-2 and F3-3 (Jepson et al. 1994, O'Connel et al. 1988, Timmermann 1989, Irzyk and Fuerst 1997, Rossini et al. 1996), but also by the tau ZmGSTUI-1 (Irzik and Fuerst 1997). GSTs involved in the metabolism of dimethenamid and metolachlor in wheat and soybean, respectively, are tau class (Riechers et al. 1997, Skipsey et al. 1997).

On the whole a very large number of plant GST enzymes/ genes have been identified in the last years, many of them as a result of sequencing projects. It now appears likely that most plant species possess up to 25 or more different Gst genes. Table 5 shows a few examples, with the relative abun- dance of the different classes, when known.

Although GS-conjugates of herbicides are usually much less toxic than the original compound, their accumulation in the cytosol could still be harmful, either because, as potential inhibitors of GSTs, they may prevent conjugation of addi- tional xenobiotics, or because they might be further metabolized to toxic compounds by cytosolic enzymes. Thus removal of conjugates from the cytosol and their compartmentation and immobilization in the cell is usually required for efficient detoxification.

The vacuole has long been supposed to be a primary deposition site of GS-conjugates. The transport mechanism has recently been elucidated with the discovery, in isolated bar- ley vacuoles, of an active GS-X vacuolar pump, capable to specifically direct the transfer of a variety of GS-conjugates across the tonoplast (Martinoia et al. 1993). This pump is functionally equivalent to the GS-conjugate Mg 2§ ATPases of mammalian cells, which are coded by members of the MRP subfamily of the ABC transporter genes (Rea 1999, Sfinchez-Fernfindez et al. 2001).

The fate of GS-conjugates after vacuolar import is less defined. One possibility is that they are simply trapped as such in an organelle with limited metabolic activity to eliminate them from the cytosol, where their accumulation is potentially harmful. Such a process has been termed 'storage excretion' (Martinoia et al. 1993). However, there is evidence that GS-conjugates within the vacuole can go through further processing and degradation, and their metabolites can be recicled to the cytosol and/or definitively excreted to the apoplast. Wolf and coworkers (1996) showed that, in bar- ley, a vacuolar carboxypeptidase degrades GS-conjugates, including alachlor-GS, by cleaving the glycine residue of the GSH moiety. Also the glutamic acid residue can be succes- sively removed (Lamoureux and Rusness 1989a), possibly by a vacuolar dipeptidase (Schr6der 1997). The identification of several sulphonated chloroacetanilide metabolites in diverse crops (Lamoureoux and Rusness 1989 b, Lamoureux et al. 1991), further suggest that the cysteine conjugates are reexported from the vacuole, metabolized by cysteine S-lyases and perhaps excreted to the apoplast where they may form residues bound to lignin and cellulose.

Sulphonated chloroacetanilide metabolites have been detected in both surface and groundwaters in areas characterized by heavy use of these herbicides like the corn belt in the USA. Runoff studies from alachlor treated experimental corn fields showed that the herbicide is rapidly metabolized to its ethansulphonic acid metabolite (ESA) that, because of its water solubility, can be easily transported by surface waters (Thurman et al. 1996). How these compounds are formed in soil, whether they are microbial or reactions from plant exu- dates is not well defined yet. On the other hand data from parallel studies on hydroponically and soil grown soybeans treated with propachlor suggest that the likely major source of ESA metabolites is microbial degradation in the soil, through a metabolic pathway involving GSH conjugation as a first step, and that the metabolites detected in plants may, at least in part, be first formed in the soil and subsequently taken up by crops (Lamoureux and Rusness 1989). Since ESA metabolites of chloroacetanilides could be toxic (toxicity data are still scant), the possibility that they could be removed from the soil by plant uptake is intriguing.

A secondary pathway for chloroacetanilide metabolism initi- ates with their hydroxylation/oxidation and involves P450 enzymes. O-demethylation of metolachlor by microsomal P450 monooxygenases has been observed in corn, mung bean and sorghum (Moreland et al. 1990,1993 1995). The O-demethylated metabolite may be further conjugated with glucose (Moreland et al. 1990) and, in analogy with other glucose- conjugated herbicides (Sandermann et al. 1991, Schmitt and Sandermann 1992, Gaillard et al. 1994) translocated into el-

Table 5: Putative number and class subdivision of expressed or potentially expressed Gst genes in three representative plant species. Arabidopsis data were obtained from EST and genomic sequences present in Gene Bank and EMBL databases

~ S ~ i ~ ' : : : ; : c l a s s ' ReferenCes

putative number phi tau zeta theta Arabidopsis 40-50 12-15 25-30 2 3-4 MiT_zi L, pers. comm.

Soybean 25 4 20 1 0 McGonigle et al. 2000

Maize 42 12 28 2 0 McGonigle et al. 2000



ther the vacuole or the apoplast. The involvement of this secondary pathway for chloroacetanilides detoxification is sup- ported by recent studies with transgenic potato plants co-expressing three human P450s. These transgenic plants show a marked tolerance increase towards several herbicides, including metolachlor and acetochlor. Moreover the microsomal fraction of recombinant yeast strains expressing the same human P450s were able to metabolise the herbicides and one of the identified metolachlor metabolites was O-demethylated metolachlor (Inui et al. 2000).

4 Enhancement of Chloroacetanilide Metabolism

Since plant tolerance to chloroacetanilides is fundamentally related to their detoxification capacity via specific biochemical reactions, it is reasonable to expect that higher tolerance may be achieved by two ways. i) An increase in metabolism rate and/or in the activity of the enzymes/genes involved; and ii) de novo acquisition of metabolic capabilities through introduction and expression of novel genes. As for the first point only few cases have been reported of induction by the herbicide itself: in maize, ZmGst2 transcription was specifically enhanced by alachlor (Rossini et al. 1998) and in wheat GST 25 and 26 were induced by alachlor and metolachlor (Mauch and Dudler 1993). On the contrary, the inductive effects of several chloroacetanilide safeners on different metabolic pathways is well documented, and seems to be the major mechanism involved in the protective action of these compounds. In particular, safeners may affect one or more of the following: glutathione metabolism, herbicide- glutathione conjugation, herbicide hydroxylation/oxidation and the ensuing conjugation with glucose, GS- or glucoside- conjugate vacuolar transport (see Hatzios 1997 for a review). For instance, fenchlorim, which is routinely used to protect rice from pretilachlor injury, has been shown to increase GSH levels, through the induction of glutathione reductase (Han and Hatzios 1991), specific GST activity (Wu et al. 1996) and Gst mRNA levels (Wu et al. 1999). In several cereal crops a strong correlation between the protective action of chloroacetanilide safeners and their ability to increase GST activity has been demonstrated (Gronwald 1989, Cole et al. 1997, Marrs 1996). In maize, the induction of expression of specific Gst genes was the most common response to safeners observed (Wiegand et al. 1986, Jepson et al. 1994, Irzik and Fuerst 1997).

The second way can be the result of different approaches, i.e. interspecific crosses, selection of spontaneous or induced mutants, genetic manipulation via recombinant DNA technology. Of these the most promising to da te is the third. Depending on the mode of action of the different classes of herbicides and the mechanisms of tolerance, three main routes have been suggested or applied to engineer tolerance in the plants of interest:

1) Overexpression of the target enzyme, such as, for instance, EPSPS in the case of glyphosate (Shah et al. 1986);

2) modification of the target enzyme, such as ALS in the case of sulphonylureas (Mazur and Falco 1989);

3) introduction of herbicide detoxifying enzymes from diverse origin. Nice examples of this is the obtainment of

metolachlor tolerant transgenic tobacco plants expressing the maize Gstl and Gst2 genes (Jepson et al. 1997) and tolerant transgenic potato plants expressing human P450 genes.

5 (2,4-Dichlorophenoxy)acetic acid;2,4-D

The phenoxyalkanoic (or aryloxyalkanoic) acids are a family of selective herbicides that were developed in the 1940s. Some members of this group notably the phenoxyacetic acids like 2,4-dichlorophenoxyacetic acid (2,4-D) were among the first commercially successful synthetic herbicides. 2,4-D, is still one of the most widely applied herbicide throughout the world for both crop and non-crop uses (Fig. 5, Table 6).

2,4-D is a systemic hormone-type herbicide that acts as an antagonist to natural plant growth regulators causing ab- normal growth and development in susceptible dicotyledonous plants.

O--CH2-COOH

CI

Fig. 5: 2,4-dichlorophenoxyacetic acid (2,4-D)

Table 6: Physical properties of 2,4-D

Molecular formula C8HECI203


Water solubility 45,000mg/L (Na salt 20~ 890 (free acid at 25oC)

Vapour pressure 1 x 10 -s Pa (20~

Partition coefficients Log Kow = 2.58-2.53 (pH 1)

The wide use of 2,4-D and fears that it might be a danger to human health has prompted extensive research and evalua- tion of its environmental toxicology and biodegradation. The acid form (2,4-dichlorophenoxy)acetic acid is the herbicidal moiety. However most commercial formulations contain 2,4- D as an inorganic or amine salt or as an ester. Results from a number of studies have indicated that the rates of breakdown of the amine and ester forms in soil are similar (Wilson et al. 1997). In soils the 2,4-D ester underwent hydrolysis to the acid at a rate dependent on pH. For example the half- life at pH 8.0 for the butoxy ester was 9 hours, whereas at pH 5.0 it was more than one year (Smith 1988).

5.1 2,4-D metabolism in soils

In the soil 2,4-D is relatively immobile (lowest average depth detected 60 cm.) and has short half-life. In studies in the



USA, its average half-life in soils ranged from 6.4 days in southern US soils to 8.3 days in soils with high organic matter (Roberts 1998). The half-life in natural water was two to four weeks, although in areas such as a treated rice paddy, the half-life was as short as one day. The acid form of 2,4- D, as well as the amine and ester chemical groups, is metabolised by soil microorganisms. Microbial degradation is brought about by a variety of microorganisms and soils pre- viously treated with 2,4-D may become enriched in these microorganisms and degradation may take place without a significant lag period. Typically microbial metabolism leads either to the formation of Krebs cycle acids or by complete degradation to carbon dioxide. As a result some of the carbon will become associated with the microbial biomass.

Details of the metabolism of 2,4-D by soil microbes are be- yond the scope of this review. In general the main steps in the pathways of microbial degradation are the enzymatic cleavage of the side chain forming glyoxylate and the parent phe- nol, followed by oxidation and cleavage of the aromatic ring.

5.2 2,4-D metabolism in plants

2,4-D is taken up, systemically translocated and actively metabolised in several dicotyledonous and monocotyle- donous plant species, for example bean, soybean and maize and wheat respectively. In contrast to the degradative metabolism of soil microbes, the major metabolic pathways in plants involve hydroxylation and conjugation with sugars or amino acids. The principal pathways are shown in Fig. 6.

The pathways used and the mixture of products formed depend on the plant species. In addition some pathways are complicated further by steps that include migration of the chlorine substituent to another site along with hydroxylation of the aromatic ring. For example in bean and soybean 2,3 dichloro-4-phenoxyacetic acid is a major product of metabolism. In several species the hydroxylated products are con-

0 II O-CH2-C--NH--asp ~ CI ql

CI

O-CH2-COOH @c, OH

CI

l, O--CH2-CH2-CH2-COOH @c, CI

O--CH2-COOH

CI o--glucose

O-CH2-COOH

2 I, ~ C I

CI OH

Fig. 6: Metabolism of 2,4-D. Reactions shown are 1. Hydroxylation, 2. Hydroxylation and chlorine migration, 3. Glucose conjugation, 4. Side chain extension and 5. Amino acid conjugation

jugated to form O-~-D-glucosides. Conjugates of aspartate and glutamate are the most frequently observed amino acid conjugates. Formation of these conjugates may be followed by subcellular compartmentation in the vacuole. The ]]-D- glucoside conjugate was the first reported example of vacuolar sequestration of a herbicidal conjugate (Schmitt and Sandermann 1982). Recent studies suggest that the glucosyl conjugation takes place in the cytosol and members of ATP- binding cassette (ABe) transporter family catalyses the transport of the glucoside into the vacuole. In addition to the hydroxylation and conjugation reactions, some grasses can metabolise 2,4-D by lengthening of the side chain. For example in alfalfa (Medicago sativa L.) side chain extension forming 2,4-dichlorobutyric acid is serves as resistant mechanism to 2,4-D (Roberts 1998).

At some sites 2,4-D has been identified as an important pollutant. However under normal conditions contamination from the agricultural use of the herbicide does not lead to persistence in soil, water and vegetation as both plants and microorganisms can metabolise 2,4-D. Whereas microorganisms can effect complete biodegradation, metabolism of 2,4- D in plants leads to chemical transformation of the compound followed by sequestration of the non-toxic products. Remediation of sites affected for example by point source contamination of the herbicide will require the concerted activities of soil microorganisms (bioremediation) and plants (phytoremediation). Harvesting of the plants, followed by a post-harvest polishing step such as composting could bring about complete degradation of the parent herbicide. In addition the genes encoding the enzymes that catalyse degradation in microbes are potential resources for enhancement of the biodegradation of the herbicide.

6 Conclusions

Persistent herbicides from four compound classes have metabolism in plants via distinct pathways. All of them follow patterns that have been well described during recent years. The metabolic net includes the well known three phases cas- cade (Coupland 1991) and compounds are detoxified via conjugation reactions. Fig. 7 shows that upon conjugation a number ofenzymatic reactions convert the compounds into metabolites the final fate of which is either the apoplastic space and cell wall or the vacuole. The latter compartment is thought as a locus for storage excretion, and thus metabolites will become available to the food chain when the plant is eaten or dies. On the basis of current literature, it is un- likely to find total mineralization of herbicides to CO 2 in higher plants. Bound residues may, however, be unavailable to the members of the food web.

Soil applied herbicides tend to be adsorbed in soil particles and pose more problems for bioremediation. No valid con- cept exists for the solubilization of soil bound herbicides and their subsequent uptake into plants for remediation. Quite frequently, microbial metabolism is found to alter the active compounds and to partially mineralize them (Scheu- nert and Schr6der 1998). However, microbes do so only in the course of co-metabolism and leave residues behind once they get access to other energy sources.

ESPR - Environ Sci & Pollut Res 9 (1) 2002 2 5


Phase I

Phase II

Fate o f herbicides in plants

Chloroacet- Ureas Aryloxyalkanoic anilides acids

Triazines

�9 demethylatiin ' 1 hydroxylation

demethylation

I I Phase HI vacuole vacuole vacuole vacuole

soluble metabolites . . . . -1, secondary conjugates

Fig. 7" Fate of herbicides in ~lants. The thickness and length of the arrows indicates the importance of the respective metabolic step for the corresponding herbicide. Generally herbicides are activated, conjugated and further metabolized according to Coupland's three phase scheme (Coupland 1991) and finally stored as soluble non toxic compounds or bound residues

For the removal of diffuse pollution of persistent herbicides f rom agricultural or set aside land, agronomic practice has to be applied. Farming techniques will provide the correct seed density and nutrient supply, and the growing canopy will prevent the pollutant f rom leaching into ground water or surface runoff. Treatment with herbicide antidotes may strengthen the plant performance, but ways to enhance the uptake f rom soil have to be found. As some compounds are adsorbed to organic matter in the soil, setting aside the land before agronomic phytoremediat ioon programs are started seems help- ful, as this might solubilize the herbicidal pollutants for phytoextract ion. In any case, care should be taken to prevent the plants used for phytoremediat ion being used as fodder.

Wi th view to the growing demands of the wor ld popula t ion for heal thy food, clean water, space for housing and recrea- tion, phytoremedia t ion might present an ecologically and economical ly sound wa y for the removal of persistent herbicides f rom ecosystems.

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Received: December 17th, 2001 Accepted: December 20th, 2001

OnlineFirst: December 21st, 2001


Exploiting plant metabolism for the phytoremediation of persistent herbicides

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