2002 phytoremediation pesticides.pdf

P e r s i s t e n t P e s t i c i d e s P h y t o r e m e d i a t i o n

Phytoremediation: Persistent Pesticides : I ' II

Prospects and Limitations of Phytoremediation for the Removal of Persistent Pesticides in the Environment

Qasim C ha udh r y 1, Peter Schr6der 2, Daniele Werck-Reichhart 3, Wlodzimierz Grajek 4 and Roman Marecik 4

t Central Science Laboratory, Department for Environment, Food and Rural Affairs, Sand Hutton, York Y041 1LZ, United Kingdom z Institute of Soil Ecology, GSF National Research Center for Environment and Health, Neuherberg, Germany 3 Dept of Stress Response, IBMP-CNRS, Strasbourg, France 4Department of Biotechnology and Food Microbiology, A. Cieszkowski Agricultural University of Poznan, Poland

Corresponding author: Qasim Chaudhry, Central Science Laboratory, Department for Environment, Food and Rural Affairs, Sand Hutton, York Y041 1LZ, United Kingdom; e-maih q.chaudhry@r

DOI: http://dx.doi.org/10.1065/esDr2001.09.084.1

Abstract. The environmental problems that have arisen from the use of persistent pesticides in the past, and potential sources of further contamination have been discussed. The potential and limitations of phytoremediation for removal of pesticides in the environment have been reviewed. The enzymatic processes in plants that are known to be involved in phytodegradation of pesticides, and possibilities for enhancing them have also been discussed.

Keywords: Carbamate; contamination; environment; organochlorine; organophosphate; persistent pesticides; phytodegradation; phytoremediation; plant uptake

1 Background

The widespread use of synthetic organic pesticides started after the 1940s. A number of potent and broad-spectrum organochlorine (OC) and organophosphorus (OP) insecticides were developed between 1939 and 1950. These were followed by the development of carbamates and pyrethroids in the next decade. During this period, synthetic pesticides were widely used for protecting crops and controlling dis- ease vectors throughout the world, especially in developing countries in the tropical region [1]. Unfortunately, many of the early OCs and OPs were developed with little regard to problems that resulted from their large-scale use. Not only were there numerous incidents of acute toxicity in the short term [2], but also that residues of some persistent pesticides lingered on in the environment decades after abandoning their use (Table 1). Their bioaccumulation and bioconcentration in the food chain has since posed long-term risks to non-target species, including humans. Some persistent pesticides in the environment have been linked to disruption of endocrine hormones, increased risk of cancers, dysfunction of the immune system, and reproductive and developmental abnormalities in a number of aquatic and terrestrial species [7-9]. The widespread use of persistent pesticides also led to the development of new pests and diseases, and selection of resistance in a number of pest species.

By early 1970s, most uses of DDT and other persistent pesticides had been banned in many countries, although some applications continued into the mid-1980s. A limited use of some of the compounds is still allowed in many countries; for instance, DDT is used in over 20 countries in Asia and South Pacific for the control of insect vectors of tropical diseases, such as malaria. Since banning of persistent pesticides, and introduction of relatively more biodegradable ones (OPs, pyrethroids), their levels in the environment have been declining slowly [10-12]. However, remnants of their former use can still be detected in virtually every environmental compartment around the world. The United Nations Envi- ronment Programme (UNEP) has identified twelve persistent organic pollutants that require urgent regulatory atten- tion. Nine of these are pesticides (aldrin, chlordane, DDT, dieldrin, endrin, heptachlor, HCB, mirex, and toxaphene), whilst the other three include polychlorinated biphenyls (PCBs), dioxins and furans [13]. Traces of a number of pesticides (predominantly OCs) have even been detected in regions, such as the Arctic, where these chemicals were never used [14,15]. In addition to this diffuse contamination, seri- ous risks of further environmental pollution exist in the form of huge stockpiles of obsolete pesticides in many countries. The Food and Agriculture Organization (FAO) of the United Nations has estimated that at least 100,000 tons of unwanted pesticides are stored in developing countries, of which around 20,000 tons are in Africa (Fig. 1).

Similar stockpiles exist in some countries of the Central and Eastern Europe, such as Poland [16,17], Latvia [18], and Bielarus [19]. In many cases, pesticide containers had been stored in the open or in underground bunkers. The deteriora- tion of containers over time led to leakage of pesticides and contamination of soil and water in the immediate vicinity, and subsequently in areas far from the storage sites [20,21]. The unwanted pesticides mainly include both OCs (DDT, dieldrin, HCHs), and OPs (parathion, methyl-parathion, dichlorvos and monocrotophos). The FAO has estimated the cost of removing pesticide from Africa, and transporting and inciner- ating them in the EU at $3,500 to $4,000 per ton [22]. For most developing countries where obsolete pesticides are stockpiled, such costs are prohibitively high, and there is a need

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Phytoremediation Persistent Pesticides

Table 1: Persistent pesticides and their transformation products

HCB

y-HCH

DDT

DDE

OC

OC

2.7 to 22.9 years (HCB)

10 to 15 years

5.73

3.76

6.19

5.69

HCB

Endrine

Toxaphene

Mirex

OC, cyclodiene

OC, camphene

OC

Up to 12 years

100 days to 12 years

Up to 10 years

5.2

5.78-6.79

7.18

CI CI~CI

CI" y "CI Cl

~

Cl / ~"C I Cl

DDT

Cl

C'\l

Endrine

CH 2 Toxaphene

C ~ "CI r Cl CI

Mirex

Atdrin

Oielddn

OC, cyclodiene 5 years (dieldrin) 6.5

5.4

c,' \ i CH 2

Aldr in

Cl Cl

CI Cl

cl \[ CH 2

Dieledrin

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Persistent Pesticides Phytoremediation

Table 1 : Persistent pesticides and their transformation products (cont'~

PestiCides/transformation product :: Chemical class

T Heptachlor OC

Chlordane

Endosulfan

Parathion

Carbaryl

Dimethoate

Persistence [3]

OC, cyclodiene

OC

OP

Carbamate

OP

Up to 2 years

1 year

50 days

14 days

10 days

7 days

Log K=, [4,5,6]

5.47-6.10

6.00

4.70

3.83

2.81

0.74

Chemical StructUre

Cl\ Cl

CI ~ C I 0,2 Cl

Heptachlor

c,

c

cis and trans Chlordane

Endosulfan

O"

Parathion

O - - C - - ~ I - - C H 3

Carbaryl

s H3C\ II H2 H

\ o ~ P ~ / C ~ . / m ~ H ~ C ~ O / S C ~ ~"CH~

II o

Dimethoate

Low = half life <30 days; Medium = half life 30-100 days; High = half life >100 days

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Fig. 1 : Regions in Africa where obsolete pesticides have been stockpiled, of the United Nations

for low-cost, and environment-friendly methods for the safe disposal of remaining pesticide stocks, and removal of their residues from contaminated environments.

Although known for centuries, the use of plants to clean up contaminated environments has gained scientific credibility over the last few decades. The main emphasis until now has been to use plants with natural ability to hyperaccumulate inorganic pollutants, such as toxic heavy metals. The technology, however, could also offer benefits in cleaning up the environments contaminated with hazardous organic pollutants such as pesticides.

2 D i f fused C o n t a m i n a t i o n of the E n v i r o n m e n t

It is outside the scope of this review to cover the full extent of contamination of the global environment by persistent pesticides. Suffice to say that a number of man-made pesticides, such as DDT and HCHs, have now become a ubiquitous part of our environment, and that their traces are detectable in almost every environmental compartment. When released into the environment, pesticides disperse through volatilisation, leaching, run-offs and drainage, and become partitioned between different geo- and bio-phases. The range and degree of environmental contamination is, however, dependent on a number of environmental factors and physicochemical properties of a pesticide, such as volatility, chemical reactivity, absorption and adsorption, aqueous solubil- ity, distribution between polar and non-polar phases (log Kow) and between soil and water (Kd) [23]. Some pesticides, e.g. DDT and dieldrin, have been shown to be more extract- able from soil than others, such as methyl-parathion [24]. In general, pesticides with Kd greater than 3, and/or half-

reproduced with permission from the Food and Agricultural Organization (FAO)

lives of less than 3 months, may not leach into the underground water [25]. Also, unless used repeatedly, the presence of highly persistent pesticides (half-life >6 months in soil) in run-offs may not pose a sign!ficant hazard to the environment [25]. The semi-volatile pesticides, on the other hand, are dispersed over a wide range in the environment, and tend to accumulate in relatively cooler regions that may be far away from the source where they were used.

Once in the environment, the persistence of a pesticide is dependent on its chemical stability, degradability by micro-organisms and uptake by aquatic and terrestrial species including plants. Most semi-volatile pesticides tend to become concentrated in atmospheric water, and precipitate back to land and water in the form of snow, rain or fog [26,27]. The process is dependent on climatic factors and usage of pesticides in the region. The pesticides frequently detected in rainwater are the compounds that were used in the past (e.g. DDT and lindane), as well as some newer pesticides (e.g. atrazine, simazine) that are relatively more water soluble and thus more mobile in the environment [28,29]. The most frequently detected pesticides in rainwater in Europe were HCHs, followed by atrazine, MCPA, simazine, dichloprop, isoproturan, mecoprop, DDT, terbuthylazine and aldrin [30]. Although most pesticides detected were below 100 ng/1 (permissible limit for drinking water), some samples were found to contain concentrations in the range of microgram per litre. A study in USA showed that traces of many pesticides, mostly OPs and their oxons (diazinon, chlorpyrifos, malathion, methyl-parathion), were present in fog-water [31].

Depending on their physico-chemical properties, most pesticides used on land end up in aquatic environments. Through the volatilisation and condensation cycle, and seepage and



run-offs, pesticides get into surface and ground water, rivers, lakes, seas and oceans [32,33]. The analysis of OC pesticides in water [34,35], sediments [36], marine species [37-39] and their products such as cod liver oil [40] has been used to assess the levels and patterns of pollution of the marine environment. The distribution patterns of different OC pesticides have been reported to differ in surface and core sediments from the Gulf of Alaska, Bering Sea and Chukchi Sea. Compared to a uniform distribution of HCHs and HCB, DDT and metabolites have been reported to show a decreasing trend from north to south [36]. Similarly, the findings that concentrations of some OC pesticides in tropical fish were lower than those in fish of the temperate regions [41,42], reflect dispersion of semi- volatile compounds from tropical to temperate regions. The bioaccumulation patterns of OCs by aquatic species have also been shown to differ with the depth of water; with the deep- sea biota showing much higher burdens compared to surface- living species of the same region [43-45]. Indeed, the bio- and geo-phases of the deep-sea have been regarded the global sink for persistent semi-volatile contaminants, in a way similar to that of soil [45].

Most fungicides used in the EU region have been regarded as not persistent enough to have a significant impact on estuarine and marine environments, and their traces in drainage canals and river samples could be linked to known agricultural usage of the compounds [46]. However, the triazole fungicides, flutriafol, epoxiconazole and triadimenol, have been reported to be very persistent (around 2 year half lives at 10~ under laboratory conditions, whereas propiconazole was moderately persistent (half-life around 200 days) [47].

Pesticides undergo a number of degradation processes during storage, mixing, and during and after application. These reactions reach equilibrium after some time, and so far no data are available on the specific percentages or the effects of degradation products resulting from breakdown of pesticides. It is, nevertheless, well known that some pesticides could be degraded to still active pesticides [48]. The main transformation or degradation products of pesticides have been given in Table 2.

Table 2: Environmental breakdown products of some Pesticides

Pesticide Breakdown products [153]

DDT DDE, DDD, DDA (a)

Aldicarb AIdicarb sulfoxide, aldicarb sulfone

Carbaryl 4-hydroxy carbaryl

Diazinon Hydroxy diazinon

Aldrin Dieldrin

Fonofos Fonofoxon

Parathion Diethylthiophosphate and p-nitrophen01, aminoparathion

Malathion Malathion carboxylic acid

Carbosulfan Carbofuran

Thiocarb Methomyl

(a) DDA (2,2-bis(chlorophenyl)acetic acid), has been regarded as an important environmental breakdown product of DDT, due to its leaching into ground water [49]

3 Potential Sources of Further Contamination

Whilst banning or restricting the use of persistent pesticides has reduced further inputs into the environment, there are still substantial risks of pollution from stockpiles of obsolete pesticides in many countries. A typical example, which reflects the situation in some countries of the Central and Eastern Europe is Poland, where up to 60,000 tons of unwanted pesticides may have been stored. These stocks of pesticides are scattered throughout the country in warehouses and on individual farms, of which up to 12000 tons are stored in underground bunkers or 'tombs'. The widespread use of persistent pesticides in the 1950s, together with acci- dental releases from agricultural, industrial, transportation and marketing sectors, has already led to severe environmental problems in Poland [50]. A recent study has shown that all 59 soil samples collected in Poland during 1990- 1994 were contaminated with DDT and metabolites, PCBs and HCHs, and that only few of the samples did not contain HCBs or chlordane [51]. Added to this situation are the further risks posed by huge stocks of unwanted pesticides that have reportedly resulted from improperly controlled past system for imports, subsidies and distribution, and also due to banning and restricting the use of persistent pesticides in Poland [16,17]. According to an estimate, around 5,000- 8,000 ton pesticides passed their expiry date every year during the 1980s alone. Some of the growing stocks of pesticides in Poland were dumped in waste stores built during the 1970s. A typical dump was built of concrete rings, 1-2 meter in diameter and 1-3 meter high. The base and top lid were also made of concrete, and the internal and external surface was insulated with tar before placing it underground. With a space of I to 10 m 3, most dumps were designed to store between 1 to 10 tonnes of chemicals [16]. The exact number of such dumps in Poland is not known, as many were built without permission from the authorities. In some cases, non-insulated ground pits were used as dumps. In- deed, up to 1,000 such sites may exist around the country, with a stockpile of up to 12,000 ton of hazardous chemicals. In addition to being poorly constructed, many dumps were built at ground-water level, in areas that were periodi- cally flooded, in limestone and sandy formations, or near a source of drinking water, river, lake, or local dwellings [52].

In 1994, around 65 tons of unwanted pesticides were moved to a safer unit [53], and in 1995, 32 tons of pesticides were transferred from two sites to a new specially built store. Over 2,000 ton of highly contaminated soil were also transferred to the new site where pesticide residues were rinsed into a retention and evaporation tank to allow environmental- and microbial-degradation. The studies showed that concrete walls offered a poor barrier to pesticide diffusion from the dumps, and leakage of biologically-active substances could lead to a hazard zone that could extend to the range of a few kilometres. The main pesticides of concern in the dumps include chlorinated hydrocarbons, organomercurial and organoarsenical compounds, nitrophenols, and relatively smaller amounts of carbamates and phosphoric acid esters. According to an estimate, around 30% of the stored pesticides are OCs, of which up to 50% may be DDT [54]. On the basis of chemical analysis of samples from 128 sites, a

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Table 3: The chemicals most frequently found in pesticide storage bunkers in Poland [55]

Organochlorines 29-30

Inorganics 15-17

Dithiocarbamates 12-14

Phenoxyacids 12-15

Organophosphorous 8-10

Nitro-derivatives 8

Carbamates 4

Tar 3-6

Quaternary amines 2

Phthalimides 1

Table 4: The organochlorine compounds most frequently found in pesticide storage bunkers in Poland [55]

DDT 42

Chlorinated camphenes 20

DMDT 11

HCH and lindane 10

Diene derivatives 4

Hexachlorobenzene, tetradifon, 13 kelevan, ovotran, endosulfan

list of the 10 most frequently found chemical groups in the dumps is presented in Tables 3 and 4 [55]. Around 75% of the stored pesticides are in solid form, with the remaining in liquid formulations. More than half of the sites have not maintained a complete, or any, record of the quantity or the nature of stored chemicals, and efforts have recently been made to develop a computerised inventory of the sites and the stored pesticides [56].

There are also reports to indicate that similar problems re- garding pesticide wastes exist in other neighbouring countries, such as Latvia [18], and Bielarus [19], where DDT and lindane were among the most widely used pesticides in the past decades. It has been estimated that the unwanted pesticides stored in Latvia occupy 1400 m 3 space at around 120 underground dumps [18].

There is an urgent need for strategies for the safe disposal of remaining stockpiles of persistent pesticides in developing and developed countries, and for remedial actions to decon- taminate polluted soil and groundwater to minimise the risk of further environmental pollution. However, compared to the disposal of unwanted stocks, the decontamination of polluted environments may be a very expensive and diffi- cult task. In fact, most physical and chemical methods for pesticide decontamination may not be feasible or practicable in a field situation, and this demands a thorough assessment

of the biological methods that are available for in situ remediation of pesticide polluted environments.

4 Available Options for Decontamination of Pesticides

A number of methods are available for the deactivation of undiluted and diluted pesticides in left over formulations and wastewaters. Depending on the chemical nature of a pesticide, treatment with a chemical reagent such as a strong acid or alkali, oxidants such as hydrogen peroxide or ozone, or catalytic reduction is usually sufficient to cause deactivation due to the cleavage of one or more bonds in the active ingredient. The chemical oxidation processes currently used for pesticide degradation in wastewater include heteroge- neous TiO z photocatalysis, ozonation or treatment with Fenton's reagent (Fe § and H202) [57,58]. For pesticide-contaminated soil or groundwater, most of the chemical methods may be prohibitively expensive or even impracticable, and the resulting products may not always be safer than the parent compounds.

The other main option for disposal of unwanted pesticides involves degradation by thermal processes, e.g. high-temperature incineration, or incineration in cement kilns. However, initial trials using cement kilns in Pakistan, where up to 5000 tons of unwanted pesticides may be stored, and other countries, have highlighted operational difficulties and problems related to the control of emissions [59]. Thermal desorption of pesticides bound to soil has been reported to be an effective technique for decontamination of OC pesticides. At soil treatment temperature between 260 to 538 ~ the thermal desorption of pesticides to 0.1-1.0 mg/kg levels could be achieved within 8-40 min, with removal efficiencies of >99%. At treatment temperatures of <277~ the removal efficacy of DDT was affected due to its breakdown to DDE [60]. Solar treatment of soil under plastic sheets has also been reported to enhance degradation of the OP pesticide quinalphos [61]. Guidelines for the prevention and disposal of obsolete pesticide stocks have also been published by the Food and Agricul- ture Organization of the United Nations [62].

In some countries, large quantities of unwanted pesticides were mixed with lime and buried underground during the 1970s. The soil samples from two of such sites in Greece have recently been analysed [63]. The results showed no detectable residues of the OCs aldrin, dieldrin, lindane, p,p'- DDT, or the OPs dimethoate, ethyl- and methyl parathion and diazinon, with only traces of ethyl-parathion detectable in some samples [63]. Although disappearance of the pesticides was attributed to degradation by lime at higher pH, the analyses did not include identification of the resulting degradation products, or the levels of pesticides or their breakdown products in groundwater.

The two main biological methods for the removal and/or decomposition of pesticides include bioremediation using micro-organisms [64] and phytoremediation using plants [65- 67]. Over the years, both fields have emerged as low-cost, environment-friendly technologies, backed by well studied science. Certain enzyme preparations have also been reported



to be useful in degrading pesticides. For example, an aqueous fire-fighting foam containing OP-hydrolase has been reported to effectively degrade a number of OP compounds [68].

5 Phytoremediation

The ability of certain plants to remove pollutant chemicals from the environment is well known, and some species have been well studied for their ability to 'hyperaccumulate' inorganic pollutants such as toxic heavy metals [69-71]. Some plants have also been shown to absorb organic pollutants from contaminated soil, water and air through root mass and leaf cuticle, which may be translocated to other parts of the plant for enzymatic degradation or storage [72-74].

6 Uptake of Pesticides by Plants

The uptake of pesticides by plants is dependent on physicochemical properties of the compounds, mode of application, soil type, climatic factors, and plant species. The compounds absorbed through plant roots may be translocated to other parts via xylem. The permeation from plant roots to xylem is, however, optimal for those compounds that are only slightly hydrophobic (log Kow 1.8) [75]. More hydrophobic compounds tend to bind with lipid membranes in the roots. Thus xylem translocation of compounds with log Kow >3 may be severely limited in plants [76-78]. The hydrophilic compounds, on the other hand, have a limited absorption through the leaf cuticle waxes. The microbial activity in the rhizosphere also plays a crucial role in transforming pollutant chemicals [79], which may aid root uptake and further degrada6on in plants. The microbial transformations in the rhizosphere, therefore, must be considered an integral com- ponent of phytoremediation.

A number of studies have attempted to determine the absorption and translocation of persistent pesticides in plants. For example, it has been reported that rape seedlings grown from seeds treated with lindane (7-HCH) [80], or maize plants grown in lindane-treated soil [81] absorbed and translocated the pesticide. The absorption of HCB by plant root and leaf, but lack of translocation, has been demonstrated [82], whilst 14C-labelled HCB sprayed on Norway spruce stem cuttings was translocated to the needles [83]. For semi-volatile pesticides, the form of application also plays a crucial role in the uptake by plants. The application of 14C-lindane along with nutrients to coffee plants resulted in absorption and accumulation of the pesticide in plant roots. Compared to this,, almost 90% of the pesticide was lost due to volatilisation, when applied to the foliage of potted coffee plants [84]. Simi- larly, more than 70% of 14C-etrimfos applied to bean and corn seedlings was lost due to volatilisation [85]. Whilst dieldrin, endrin and heptachlor were absorbed from contaminated soil and translocated in soybean plants, the residues of DDT in the foliage were shown to result from vapour uptake of the pesticide [86]. Similarly, the levels of p,p'-DDT, DDE, HCB and HCH detected in foliage of the dwarf bean Phaseolus vulgaris were not dependent on pesticide concen- tration in soil. The xylem translocation of the pesticides in the bean plants was negligible, and the main route of uptake

by the foliage appeared to be through absorption of vapours [87]. The uptake of p,p'-DDT by the cowpea Vigna ungui- culata from soil treated with 14C-labelled pesticide has been reported [88}. Highly weathered soil residues of p,p'-DDE have been shown to be accumulated in the roots of lucerne, ryegrass, and pole bean, but none of the plants translocated the compound to the shoot [89].

Studies have shown a substantial translocation of heptachlor epoxide, a lesser translocation of T-chlordane, but no translocation of dieldrin in forage/pasture species from soil contaminated with aged pesticide residues [90]. The residues of dieldrin, lindane, DDT and DDE, absorbed from pesticide contaminated soil, were mainly accumulated in roots of the grass Lolium perenne, whereas only trace amounts were detectable in the shoots [91]. Similarly, the residues of DDT and aldrin absorbed from contaminated soil by sweet potato tubers were concentrated (50-100x) in the peel compared to that in the pulp [92]. A recent study has shown that weathered soil residues of chlordane were accumulated in root tissues in the 12 food crops tested, which also translocated to aerial tissues in some cases [93]. The residues of chlordane were detectable in edible root tissue in carrot, beet, and potato, and in edible aerial tissues in spinach, let- tuce, dandelion, and zucchini, trace amounts in edible parts of the bush bean and eggplant, but not in edible parts of tomatoes, peppers, and corn. The absorption of 14C- dimethoate by garden pea plants was greater than that of 14C-Malathion, which reflected the differences in aqueous solubilities and coefficients of partition (log Kow) of the two compounds [94].

The uptake and translocation of persistent pesticides has also been studied in plants that grow in aquatic environments. A study of the translocation of dieldrin, methoxy- chlor, and Mirex in red mangrove (Rhizophora mangle L.) seedlings showed higher residues of dieldrin in hypocotyl and leaves of the seedlings than the other two pesticides [95]. The uptake and accumulation of atrazine, lindane and chlordane in the rooted aquatic plant Hydrilla verticillata (Royle) have been reported, with bioconcentration factors of 9.62, 38.15 and 1060.95 for the respective pesticides [96].

The factors that increase hydration of leaf cuticle are thought to increase permeation of hydrophilic pesticides (low log Kow), whereas factors that reduce wax viscosity enhance the uptake of non-polar compounds (high log Kow) [97]. Fol- lowing foliage treatment of plants with methyl 14C-labelled fenitrothion, the pesticide was found to be stored in leaf cuticular waxes [98]. Indeed, the uptake of airborne xenobiotics by pine needles has been used as an environmental indicator of the presence of PCBs, DDT, DDE, HCH and HCB in the environment [99].

It is evident from the published data that uptake and translocation of non-ionic pesticides varies greatly between plant species, and is dependent on a number of factors, most im- portantly the physicochemical properties of individual compounds, and in particular the coefficient of partition (log Kow ). Thus the uptake and translocation of more hydrophobic pesticides, and consequently their metabolic degradation, may only be limited in plants. The transformation of

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pesticides in the environment, and also due to microbial activity in the rhizosphere, could result in products that are more efficiendy absorbed and translocated in plants. Thus the factors that enhance rhizospheric microbial activity should also be expected to increase the overall efficacy of pesticide phytoremediation.

7 Degradation of Pesticides by Plants

Unlike animals, most transformation products of pesticides in plants are not destined for excretion, but instead remain in plant tissues either in conjugated soluble forms, or in insoluble bound forms. Although mineralisation (complete decomposition) of pesticides is the desired end point in phytoremediation, usually a few transformation reactions are sufficient to drastically change their biological activity. Three main reaction types are known to drive transformation of pesticides in plants: 1. Degradative, such as oxidation, hydrolysis 2. Synthetic, e.g. conjugations 3. Rearrangements, including epoxide formation.

The most commonly observed degradation reactions involve oxidation of pesticides. Some oxidation of pesticides may take place in the environment due to photo-activated or metal-catalysed reactions with molecular oxygen. Biologi- cal oxidations of pesticides are catalysed by a number of enzymes, such as cytochromes P4s0 [100], peroxidases [101] and flavin-dependent monooxygenases [102,103]. The cytochromes P4s0 superfamily is composed of at least 215 fami- lies. These microsomal enzymes, ubiquitous in both animals and plants, play a crucial role in biosynthesis and breakdown of a wide variety of endogenous and exogenous compounds through oxygenation or hydroxylation. The enhanced levels of cytochromes P4s0 in insects have also been implicated in insecticide resistance [ 104,105], and it is inter- esting to note that the same enzymes that are responsible for synthesis of plant defence chemicals also provide mechanisms for detoxification of those phytochemicals and synthetic pesticides in insects.

The metabolism of herbicides by cytochromes P4s0 in plants has been well documented since it constitutes a major deter- minant of herbicide tolerance and selectivity, and of the evolu- tion of resistant weeds [106]. Evidence, often indirect, for their involvement in the oxidative transformation of other pesticides has only been reported in few cases. For example, transformation of aldrin to dieldrin through epoxidation has been reported in vitro in root extracts from different species of beans, and varieties of peas and maize [107]. The epoxidase activity, which was the highest in microsomal fractions, was stimu- lated by the addition of NADPH. None of the preparations, however, showed metabolism of heptachlor. The pre-treatment of pea plants with aldrin was reported to induce a higher rate of aldrin metabolism [108], but later studies on the soybean root nodule aldrin epoxidase casted doubts on the participa- tion of a cytochrome P4s0 in the reaction, and suggested the possible involvement of a peroxidase [109].

A number of early reports indicated that many insecticides acted as synergists of P4so-metabolised herbicides [110,111].

It was later demonstrated that malathion, terbufos and its oxidative metabolites formed in the soil behaved as mecha- nism-dependent inactivators (suicide substrates) of cytochromes P4s0 involved in the detoxfiication of sulfonylurea and several other classes of herbicides [112-114]. A mecha- nism was proposed in plants, which is parallel to mammalian cytochromes P4s0 catalysed desulfuration of organophosphate insecticides (P=S ---) P=O) with the release of atomic sulfur that tends to covalently bind to P4s0 polypeptide [115]. The desulfuration of malathion, methidathion, isozafos and diazinon was later detected in maize or sorghum microsomes that was characteristic of a P4s0-catalysed reaction [114, 116,117]. The cytochromes P4s0 and peroxidases in the microsomal fraction from tulip bulbs (Tulipa fosteriana) have been used in studying the metabolism of model xenobiotic compounds [118]. The metabolism of different polychlorinated biphenyl (PCB) congeners by non-photosynthetic sus- pension cultures of rose [119] also appeared to involve cytochromes P4s0, as the reaction was inhibited by inhibitors of the mixed function oxidases, but not by peroxidase inhibitors. Despite such evidence, insecticide, nematicide or fungicide detoxification by plant cytochromes P4s0 has never undergone a thorough investigation. As the number of P4s0 genes in a small plant genome like that of Arabidopsis is close to 300, and specific P4s0 isoforms from mammals and insects have been found to metabolise pesticide compounds [120-121], it seems inevitable that useful catalysts for the detoxification of pesticides will also be found in plants. With the recent availability of recombinant enzymes, this field now seems widely open for further investigations.

The oxidation of pesticides in plants is, nevertheless, not as efficient as in animals, and enzymes for oxidation of cyclodiene pesticides do not appear to be widely distributed in the plant kingdom. For example, aldrin and isodrin but not heptachlor were epoxidised in bean and pea root homogenates, whilst none of the compounds was transformed in corn root homogenates [122].

Carbamate compounds are metabolised in plants through oxidation or conjugation, although the rate of transformation has been reported to differ in different plant species [123], and for different carbamate compounds [124]. The ester group in carbaryl was not hydrolysed in bean plants, although the compound was glycosylated and also transformed to some insoluble forms [125]. Carbamate compounds with an aromatic moiety have been reported to be resistant to hydrolysis in plants [125,126].

The OP compounds may be activated by certain transformation reactions in plants, such as desulfuration, sulfur oxidation, or N-dealkylation, or deactivated by O-dealkylation and hydrolysis. Desulfuration at the thiono-sulfur apparently does not take place as readily as oxidation of thio-ether sulfur in plants [127-129]. The alkyl sulfur, on the other hand, is rapidly oxidised to sulfoxide, but only slowly to the sulfone [130]. Deactivation of OPs through hydrolysis or conjugation appears to be common in plants [127,128,131]. A small proportion of 14C-etrimphos applied to bean and corn seedlings was found to be hydrolysed to 6-ethoxy-2-ethyl- 4-pyrirnidinol and 2-ethyl-4,6-pyrimidinediol [85]. Similarly, dimethoate was shown to be hydrolysed in wheat and soy-



bean grains in vitro and in vivo with cleavage of the P-C-S bond at P-S position [132]. The hydrolysis of malathion to dimethylphosphorothionate and dimethylphosphorothiolo- thionate, and oxidation to malaoxon, followed by hydrolysis to malaoxon mono- and di-acids, has been demonstrated in wheat grains [133]. The dealkylation of substituted amide and amino groups in plants has also been demonstrated by the isolation of N-hydroxymethyl intermediates of dicroto- phos [134] and carbaryl [125].

Organochlorine pesticides are apparently quite recalcitrant to metabolism both by terrestrial plants [130] and marine phytoplankton [135]. The absorption but lack of metabolism of DDT by barley plants has been reported [136]. The metabolites of DDT detected on leaf surface of apples in- cluded traces of DDD and 4,4' dichlorobenzophenone (DCBP) [137], whereas DDD and DDE but no DCBP was detected in cottonseed [138] and in extracts of cotton leaves [139]. The reductive dechlorination of p,p'- and o,p'-DDT has also been reported in the aquatic plant Elodea canadensis, and the terrestrial plant Pueraria thunbergiana [140]. The transformation appeared to take place in a non-enantio- selective manner with o,p'- and p,p'-DDD as the only resulting products, with almost a quarter of the pesticide covalently bound in Elodea canadensis. As irradiation at 300 krads did not stop transformation of DDT, the process did not appear to be driven by microbial activity. The reductive transformation of pesticides in plants is, however, less common than oxidation [130], and may not lead to their detoxification. For example, dehydrochlorination of DDT may result in equally toxic products, such as DDE.

The synthetic routes to the detoxification of xenobiotics in animals mainly involve catalysis by the enzyme glutathione- S-transferases (GST), which binds a wide variety of electro- philic compounds to the tripeptide, glutathione. The resulting hydrophilic conjugates are broken down subsequently and excreted as mercapturic acid conjugates via the kidney [141]. The role of glutathione in the metabolism of pesticides, herbicide resistance and heavy metal detoxification in plants has been discussed by Rennenberg [142] and Lamoureux and Rusness [143]. Atrazine was the first shown herbicide to be conjugated with glutathione in plants [144]. This was followed by reports that suggested detoxification of other herbicides via glutathione conjugation [145,146]. In fact, the enhanced GST activity in some plants has been associated with resistance to herbicides [147,148]. Extracts from spruce needles have been reported to contain GST activity that catalysed conjugation of pesticides [149]. The rate of in vitro glutathione conjugation of the fungicide PCNB (quintozene) was only intermediate to low as compared to that of the herbicide fluorodifen. A high rate of PCNB conjugation and metabolism has also been reported in onion [150]. However, as pesticide transformation reactions in plants are not aimed at products that are to be excreted, most conjugation reactions seem to be limited to sugars, peptides and sulfate, in descending order of frequency. Gen- erally, the transformation reactions that expose one or more functional groups in pesticides are followed by conjugation to a sugar moiety, such as glucose. A number of glucoside conjugates of pesticides have been reported in plants [151,

152], and an association between glutathione-transfer mediated and glucosyl-dependent detoxification pathways has been proposed [152]. The glucosylation of carbaryl h a s been reported following its oxidation in bean plants [125]. EPTC, butylate, terbufos and phorate have been shown to be converted to their respective sulfoxides. It is unlikely that the sulfur stems from a former glutathione conjugation reaction and subsequent degradation [153], although EPTC can be also detoxified via a glutathione conjugate in millet and corn [154]. The isolation of an enzyme that catalysed conjugation of the DDT metabolite DDA [2,2-bis-(4- chlorophenyl)-acetic acid] to acylglucoside has been reported from a soluble fraction of cultured soybean cells [155]. The unstable enzyme had a molecular weight of around 50 KDa, with isoelectric point of around pH 4.9. A study of many plant species and cell cultures showed that O-glucosyltransferase activity, which acts on DDA and other xenobiotics, is widespread in the plant kingdom, with some species showing par- ticularly high activities of the enzyme [156]. However, on the basis of current knowledge, it has to be concluded that glutathione and glucose conjugations are not primary detoxification mechanisms for majority of persistent pesticides. They may, however, play a role in downstream metabolic reactions of the breakdown products of pesticides.

Some compounds have also been reported to be altered in- trinsically, by isomerisation and changes in chirality that influence their biological effects [153].

The information on pathways for phytodegradation of persistent pesticides is limited and patchy, and there is a need for further systematic studies to realise the full potential of plants for use in phytoremediation of pesticide contaminated environments. It is, however, evident from the published examples that useful metabolic pathways exist in most, if not all, plants that can be utilised in the degradation of persistent pesticides in the environment. It is equally important to note that, unlike animals, plants do not have any specialised organ (such as liver) for detoxification of xenobiotics, and that the transformation reactions may take place at a far slower rate in plants than in animals. These processes may, however, be enhanced by using selected plants that have higher levels of pesticide degrading enzymes, or that have been engineered to express a useful enzyme. In either case, success of the technology would depend on selection of those plants that can grow in environments heavily contaminated with organic chemicals.

8 Enhancement of Phytoremediation

In the wake of the recent surge in biotechnological sciences, it is important to identify those areas that are of direct or indirect relevance to phytoremediation of pesticides. Unlike recent concerns over genetically modified food crops, a phytoremediation approach that utilises non-food plants with enhanced ability to cleanup pesticide-polluted environments may have a beneficial effect on the public perception of biotechnology. Through selective manipulation of a plant's ge- netic makeup, the capability to remove and/or degrade pesticides may be further enhanced in a number of ways. For example, a pesticide-degrading enzyme from another spe-



cies (such as a bacterium or insect) may be engineered in a plant; or the expression of an already present enzyme may be enhanced. Certain enzymes in insecticide-resistant insects, e.g. esterases and cytochromes P4s0, are known to confer enhanced detoxification of insecticides [157]. The genes that code for some of the enzymes have been the focus of studies for incorporation in bacteria and plants for subsequent use in bio/phytoremediation. For example, a cytochrome P4s0 (CYP9A1), associated with insecticide resistance in Heliotbis virescens, has been engineered in bacteria, tobacco and baculovirus for evaluation in remediation of pesticides [158]. A mammalian cytochrome P4s0 (2El), known to oxidise a wide range of halogenated hydrocarbons including trichloroethylene, has been expressed in tobacco plants. This has been shown to enhance the metabolism of trichloroethylene to up to 640 times in the transgenic tobacco plants, and also to increase the uptake and debrominat ion of ethylene dibromide [159].

The uptake and retention of organic pollutants by plants is another major limiting factor in phytoremediation of contaminated soil and water. Whilst the main focus of research has been to select plants with ability to 'hyperaccumulate' or biodegrade pollutant chemicals, it should also be possible to enhance a plant's ability to absorb and bind a specific pesticide by incorporating a gene for an anti-pesticide antibody fragment. The single chain variable fragment (scFv) of immunoglobulins contains the combining site of an antibody towards a given analyte. Through incorporation of the relevant gene, this 'subunit-antibody' can be expressed in plants to enhance their capability to bioaccumulate large quantities of the compound. Through appropriate design of immunogens, it may even be possible to engineer a group- specific antibody fragment that binds to a number of structurally similar chemicals. Libraries of large repertoires of cloned scFv antibody genes that have been constructed to 'display' antibody combining sites on the surface of recombinant phage, provide a means for rapid screening of a given analyte [160]. A number of reports have indicated successful incorporation of scFv antibody genes in plants. How- ever, in most cases, the antibody fragments used were reactive to a protein or peptide, and were aimed at developing plants that are resistant to pests and diseases, such as nema- todes [160,161], and plant pathogenic viruses [162,163]. It is only recently that the applicability of this technique to small-molecule organic chemicals (haptens) has been appre- ciated. To date, there have been only two reports on the expression of anti-atrazine and anti-paraquat antibody fragments in transgenic tobacco plants [164,165]. The anti-atrazine antibody expressing plants have been reported to effectively bind the herbicide in soil/water [166]. These few but successful examples demonstrate that functional scabs against organic chemicals can be expressed in transgenic plants. In the past decade, there have been great technological leaps in the fields of designing immunogens for haptens [167], screening scAb gene libraries [168], and expressing functional proteins in transgenic plants. By combining ex- pertise in these fields, it would be possible to develop plants with the capability to bind large quantities of a specific pesticide, or a group of pesticides, from a polluted environment. The development of 'generic' antibodies reactive to a

number of OP pesticides has already been demonstrated for analytical applications [169]. A similar approach could also pave the way for development of plants that express binding sites for several structurally-similar pesticides.

Other strategies to enhance phytoremediation ability in plants have been directed at increasing the root mass, which could also increase the uptake of certain pollutant chemicals. Some strains of Agrobacterium tumefaciens, a bacte- rial parasite of plants, are known to induce root prolifera- tion in plants. This has also been reported to enhance the uptake of PCBs and other pollutants by the infected plants from soil and groundwater, possibly due to an increased rate of cell division and root mass [170]. The newly emerging plant 'hairy root' technology, which has potential applications in secretion of natural products and recombinant proteins, may also be very useful in removing pollutants from pesticide-contaminated water [171].

9 Limitations of Phytoremediation

As discussed above, phytoremediation offers a low-cost, low maintenance, environment-friendly and renewable resource for the remediation of contaminated environments. In some situations, for example pesticide-laden soil or ground water, phytoremediation may be the only practical and economical in situ technique that can be used to remove pollutant chemicals. However, phytoremediation has its own limitations. The uptake and translocation of organic pollutants from soil through the plant root may be limited to compounds that are not highly hydrophobic, and through foliage to compounds that are not hydrophilic. Also, the technology would prob- ably have little impact in situations where low levels of a pesticide were widely distributed in the environment. Even when large amounts of pesticides were present in a given environmental compartment, such as soil or water, phytoremediation could at best be useful as a long-term strategy. This is because plant growth is dependent on a number of environmental and climatic factors, availability of water and nutrients, soil type and pH, etc. Even under the most favourable conditions, plant growth and removal/degradation of pesticides may not ex- ceed a certain rate. These limitations require integration of phytoremediation with other more immediate cleanup options, as well as more economical utilisation of the biomass produced for non-food purposes, for example energy generation. However, the role of vegetation and plant biomass as an ab- sorbent and a long-term sink for diffused pesticides must not be underestimated. There is also a need for finding ways to enhance the absorption and degradation of pesticides in plants. As microbial activity in the rhizosphere is known to aid the release of bound pesticide residues in soil, which can enhance uptake and transformation by plants, a combination of microbial bioremediation and phytoremediation is likely to be more successful in the field. Whilst the advancement in biotechnological sciences has revolutionised a number of areas of conventional science, this has had a little impact on phytoremediation until now. The amount of knowledge generated in plant biotechnology in recent years is enormous, and there are many more yet unexplored avenues that can be pursued to engineer or enhance a particular pathway for the phytodegradation of persistent pesticides in the environment.

ESPR - Environ Sci & PoIlut Res 9 (1) 2002 13


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Received: April 9th, 2001 Accepted: August 22nd, 2001

OnlineFirst: September 5th, 2001


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