degradacion de organofosforados

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Microbial degradation of organophosphorus compounds

Brajesh K. Singh1 & Allan Walker2

1Environmental Sciences, Macaulay Institute, Craigiebuckler, Aberdeen and 2Horticulture Research International, Wellesbourne, Warwick, UK

Correspondence:Brajesh Singh,

Environmental Sciences, Macaulay Institute,

Craigiebuckler, Aberdeen, AB15 8QH, UK.

Tel.: 144 1224 498200; fax: 44 1224

498207; e-mail: [email protected]

Received 16 June 2005; revised 24 November

2005; accepted 6 January 2006.

First published online April 2006.

doi:10.1111/j.1574-6976.2006.00018.x

Editor: Alexander Boronin

Keywords

organophosphorus compounds; microbial

degradation; metabolic pathways; detoxifying

enzymes; genetic basis; biotechnological

aspects.

Abstract

Synthetic organophosphorus compounds are used as pesticides, plasticizers, air

fuel ingredients and chemical warfare agents. Organophosphorus compounds are

the most widely used insecticides, accounting for an estimated 34% of world-wide

insecticide sales. Contamination of soil from pesticides as a result of their bulk

handling at the farmyard or following application in the field or accidental release

may lead occasionally to contamination of surface and ground water. Several

reports suggest that a wide range of water and terrestrial ecosystems may be

contaminated with organophosphorus compounds. These compounds possess

high mammalian toxicity and it is therefore essential to remove them from the

environments. In addition, about 200 000 metric tons of nerve (chemical warfare)agents have to be destroyed world-wide under Chemical Weapons Convention

(1993). Bioremediation can offer an efficient and cheap option for decontamina-

tion of polluted ecosystems and destruction of nerve agents. The first micro-

organism that could degrade organophosphorus compounds was isolated in 1973

and identified as Flavobacteriumsp. Since then several bacterial and a few fungal

species have been isolated which can degrade a wide range of organophosphorus

compounds in liquid cultures and soil systems. The biochemistry of organopho-

sphorus compound degradation by most of the bacteria seems to be identical, in

which a structurally similar enzyme called organophosphate hydrolase or phos-

photriesterase catalyzes the first step of the degradation. organophosphate hydro-

lase encoding geneopd(organophosphate degrading) gene has been isolated from

geographically different regions and taxonomically different species. This gene has

been sequenced, cloned in different organisms, and altered for better activity and

stability. Recently, genes with similar function but different sequences have also

been isolated and characterized. Engineered microorganisms have been tested for

their ability to degrade different organophosphorus pollutants, including nerve

agents. In this article, we review and propose pathways for degradation of some

organophosphorus compounds by microorganisms. Isolation, characterization,

utilization and manipulation of the major detoxifying enzymes and the molecular

basis of degradation are discussed. The major achievements and technological

advancements towards bioremediation of organophosphorus compounds, limita-

tions of available technologies and future challenge are also discussed.

Introduction

The excessive use of natural resources and large scale

synthesis of xenobiotic compounds have generated a num-

ber of environmental problems such as contamination of air,

water and terrestrial ecosystems, harmful effects on different

biota, and disruption of biogeochemical cycling. At the

present time, the most widely used pesticides belong to the

organophosphorus group. The first organophosphorus in-

secticide, tetraethyl pyrophosphate, was developed and used

in 1937 (Dragun et al., 1984). At the same time, twochemical warfare agents (also called nerve agents), Tabun

and Sarin, were developed and produced. Later, several

other organophosphorus pesticides were developed and

commercialized. These pesticides are widely used world-

wide to control agricultural and household pests. Overall,

organophosphorus compounds account for38% of total

pesticides used globally (Post, 1998). In the USA alone over

40 million kilos of organophosphorus are applied annually

(Mulchandani et al., 1999a; EPA, 2004). Glyphosate and

FEMS Microbiol Rev30(2006) 428471c2006 Federation of European Microbiological SocietiesPublished by Blackwell Publishing Ltd. All rights reserved


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chlorpyrifos are the most widely used in the US and account

for 20% and 11% of total pesticide use, respectively (EPA,

2004). Organophosphorus compound poisoning is a world-

wide health problem with around 3 million poisonings

and 200 000 deaths annually (Karalliedde & Senanayake,

1999; Sogorb et al., 2004). The compounds have been

implicated in several nerve and muscular diseases in humanbeings. Their acute adverse effects have been discussed by

Colborn et al. (1996) and Ragnarsdottir (2000). Immuno-

toxicity of organophosphorus compounds towards human

beings and wild-life has been reviewed by Galloway &

Handy (2003).

Continuous and excessive use of organophosphorus

compounds has led to the contamination of several ecosys-

tems in different parts of the world (EPA, 1995; McConnell

et al., 1999; Cisar & Snyder, 2000; Tse et al., 2004). For an

example, surveys revealed that 100% of sampled catchments

in Scotland and 75% of sampled aquatic sites in Wales were

contaminated with organophosphorus compounds used in

sheep dips (Boucardet al., 2004). Several organophosphorus

compounds are used on animals for the control of body

pests as several of them are fat soluble and can thus enter the

body readily through the skin and potentially find their way

into meat and milk (MAFF/HSE, 1995). Contamination of

grains, vegetables and fruits with organophosphorus com-

pounds is also well documented (Pesticide Trust 1996;

National Consumer Council 1998). Another potential and

more dangerous source of organophosphorus contamina-

tion comes from chemical warfare agents. About 200000

tons of extremely toxic organophosphorus chemical warfare

agents such as Sarin, Soman, and VX were manufactured

and are stored. As required by the Chemical WeaponConvention (CWC) 1993, these stocks must be destroyed

within 10 years of ratification by the member states. Use of

micro-organisms in detoxification decontamination of or-

ganophosphorus compounds is considered a viable and

environment friendly approach.

The available literature on the microbial degradation of

xenobiotics indicates that most studies have considered

three aspects:

(1) The fundamental basis of biodegradation.

(2) Evolution and transfer of such activities among micro-

organisms.

(3) Bioremediation techniques to detoxify contaminatedenvironments (Singhet al., 1999).

However, the use of micro-organisms for bioremediation

requires an understanding of all physiological, microbiolo-

gical, ecological, biochemical and molecular aspects in-

volved in pollutant transformation (Iranzo et al., 2001).

There are two types of xenobiotics that cause environ-

mental concerns: (1) compounds that are persistent and

therefore provide long exposure to non-target organisms

such as lindane and DDT, and (2) compounds that are

biodegradable but mobile in soil and are toxic and therefore

have the potential to pollute ground water, such as carbo-

furan. Extensive and repeated use of the same pesticide

without any crop or pesticide rotation for a number of years

has occasionally resulted in unexpected failures to control

the target organisms. It has been demonstrated that a

fraction of the soil biota can develop the ability rapidly todegrade certain soil-applied pesticides. This phenomenon

has been described as enhanced or accelerated biodegrada-

tion (Walker & Suett, 1986). The first evidence of biodegra-

dation of pesticide affecting its efficacy was reported in 1971

(Sethunathan, 1971). However, it was not until the early to

mid 1980s that the wider implication of enhanced bio-

degradation became observable in the field (Walker & Suett,

1986) and since then this phenomenon has been reported

for several other pesticides such as isofenphos (Chapman

et al., 1986), fenamiphos (Stiriling et al., 1992) and etho-

prophos (Karpouzaset al., 1999).

The practical significance of enhanced bio-degradation

depends on a number of interactive factors like the use of the

pesticides (soil or foliage applied), the frequency of use, the

interval between successive applications and the stability of

the active microflora without the presence of pesticides

(Kaufman et al., 1985). Recently, soil pH has been impli-

cated as a factor in enhanced degradation of atrazine in

different soils (Houotet al., 2000). This hypothesis has been

supported by recent reports of high enzymatic activity

(Acosta-Martinez & Tabatabai, 2000) and higher bacterial

activity at higher soil pH (Vidali, 2001). Sims et al. (2002)

suggested that soil pH may influence the rate of degradation

by affecting the uptake of the herbicide by soil micro-

organisms. The problem of enhanced bio-degradation be-came more acute, following the observation that a pesticide

can be degraded rapidly in soil from a field to which it had

never been applied before but which had been exposed to a

pesticide from the same chemical group (Prakash et al.,

1996). This phenomenon is known as cross-adaptation.

Cross-adaptation of enhanced biodegradation has been

reported within many groups of pesticide, such as the

carbamates (Morel-Chevillet et al., 1996), dicarboximides

(Mitchell & Cain, 1996) and isothiocyanates (Wartonet al.,

2002). On the other hand, only limited cross-adaptation for

enhanced biodegradation within the organophosphorus

class has been reported (Racke & Coats, 1988; Singh et al.,2005). Cross-adaptation within groups is unpredictable and

may occur only in one direction. The positive side of this

problem is that micro-organisms isolated for degradation of

one compound can be used for bioremediation of other

compounds for which no known degrading microbial

system is known. This aspect is well established for organo-

phosphorus compounds where a parathion-degrading bac-

terium was able to degrade a wide range of other structurally

similar compounds including chemical warfare agents.

FEMS Microbiol Rev30(2006) 428471 c2006 Federation of European Microbiological SocietiesPublished by Blackwell Publishing Ltd. All rights reserved

429Microbial degradation of organophosphorus compounds


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Isolation of pesticide degrading microorganisms is impor-

tant for three main reasons:

(1) To determine the mechanism of the intrinsic process of

microbial metabolism.

(2) To understand the mechanisms of gene/enzyme evolu-

tion.

(3) To use these microbes for the detoxification and decon-tamination of polluted aquatic and terrestrial environ-

ments (bioremediation).

Several microorganisms have been isolated which are able

to utilize pesticides as a source of energy. There are some

examples of fungi including Trametes hirsutus, Phanero-

chaete chrysosporium, Phanerochaete sordia and Cyathus

bullerithat are able to degrade lindane and other pesticides

(Singh & Kuhad, 1999, 2000; Singh et al., 1999). However,

most evidence suggests that soil bacteria are the principal

components responsible for enhanced bio-degradation

(Walker & Roberts, 1993). Several pure bacterial isolates

with the ability to use specific pesticides as a sole source of

carbon, nitrogen or phosphorus have been isolated (Singh

et al., 1999, 2000).

On numerous occasions, mixed bacterial cultures with

pesticide degradation ability are isolated but their individual

components are unable to utilize the chemical as an energy

source when purified (Shelton & Somich, 1988; Mandel-

baumet al., 1993; De Souzaet al., 1993; Robertset al., 1993);

an example is the organophosphorus nematicide fenami-

phos (Ou & Thomas, 1994; Singh et al., 2003b). Several

other studies failed to obtain micro-organisms capable of

growing on specific chemicals. However, this failure does

not exclude biological involvement in degradation and

could be attributed to the selection and composition of theliquid media under artificial environments, strains requiring

special growth factors, or a major role of non-culturable

microorganisms (Walker & Roberts, 1993). A recent report

of growing previously non-culturable bacteria in the labora-

tory with a simulated natural environment (Kaeberlein

et al., 2002) may lead to isolation and characterization of

several new chemical-degrading bacteria.

The main aim of this article is to review the metabolic

pathways involved in organophosphorus compound degrada-

tion. Our understanding of the molecular basis of organopho-

sphorus degradation has progressed dramatically in recent

years. Additional information has become available by gen-ome sequencing of several microorganisms and advancement

in molecular techniques. There is growing interest in devel-

oping biotechnological methods for clean up of contaminated

water and soil with organophosphorus compounds and to aid

in the destruction of large amounts of nerve agents. In this

article we also critically review recent biotechnological ad-

vancements in the development of bio-catalysts and bio-

sensors for organophosphorus compounds and their possible

application in bioremediation of contaminated ecosystems.

Chemistry and toxicology oforganophosphorus compounds

Most organophosphorus compounds are ester or thiol

derivatives of phosphoric, phosphonic or phosphoramidic

acid. Their general formula is presented in Fig. 1. R1and R2are mainly the aryl or alkyl group, which can be directly

attached to a phosphorus atom (phosphinates) or via

oxygen (phosphates) or a sulphur atom (phosphothioates).

In some cases, R1is directly bonded with phosphorus and R2with an oxygen or sulfur atom (phosphonates or thion

phosphonates, respectively). At least one of these two groups

is attached with un-, mono- or di-substituted amino groups

in phosphoramidates. The X group can be diverse and may

belong to a wide range of aliphatic, aromatic or heterocyclic

groups. The X group is also known as a leaving group

because on hydrolysis of the ester bond it is released from

phosphorus (Fig. 1) (Sogorb & Vilanova, 2002).

The mode of action of organophosphorus compounds

includes inhibition of neurotransmitter acetylcholine break-down. Acetylcholine is required for the transmission of

nerve impulses in the brain, skeletal muscles and other areas

(Toole & Toole, 1995). However, after the transmission of

the impulse, the acetylcholine must be hydrolyzed to avoid

overstimulating or overwhelming the nervous system. This

breakdown of the acetylcholine is catalyzed by an enzyme

called acetylcholine esterase. Acetylcholine esterase converts

acetylcholine into choline and acetyl CoA by binding the

substrate at its active site at serine 203 to form an enzyme

substrate complex. Further reactions involve release of cho-

line from the complex and then rapid reaction of acylated

enzymes with water to produce acetic acid and the

Fig. 1. General formula of organophosphorus compounds and major

pathway of degradation.


430 B.K. Singh & A. Walker


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regenerated acetylcholine esterase. It has been estimated that

one enzyme can hydrolyze 300 000 molecules of acetylcho-

line every minute (Ragnarsdottir, 2000).

Organophosphorus compounds inhibit the normal activ-

ity of the acetylcholine esterase by covalent bonding to the

enzyme, thereby changing its structure and function. They

bind to the serine 203 amino acid active site of acetylcholineesterase. The leaving group binds to the positive hydrogen of

His 447 and breaks off the phosphate, leaving the enzyme

phosphorylated. The regeneration of phosphorylated acet-

ylcholine esterase is very slow and may take hours or days,

resulting in accumulation of acetylcholine at the synapses.

Nerves are then overstimulated and jammed (Manahan,

1992). This inhibition causes convulsion, paralysis and

finally death for insects and mammals (Ragnarsdottir,

2000).

Microbial degradation of

organophosphorus compoundsUse of organochlorine pesticides such as dichloro-diphenyl-

trichloroethane (DDT), lindane, etc., has been reduced

drastically in developed countries due to their long persis-

tence, tendency towards bioaccumulation and potential

toxicity towards non-target organisms. This group of com-

pounds has been replaced by the less persistent and more

effective organophosphorus compounds. However, most of

the organophosphorus compounds possess high mamma-

lian toxicity. Among the organophosphorus compounds,

glyphosate, chlorpyrifos, parathion, methyl parathion, dia-

zinon, coumaphos, monocrotophos, fenamiphos and pho-

rate have been used extensively and their efficacy andenvironmental fate have been studied in detail. The chemical

and physical properties of some of these compounds are

listed in Table 1. The phosphorus is usually present either as

a phosphate ester or as a phosphonate. Being esters they

have many sites which are vulnerable to hydrolysis. The

principal reactions involved are hydrolysis, oxidation, alky-

lation and dealkylation (Singh et al., 1999). Microbial

degradation through hydrolysis of P-O-alkyl and P-O-aryl

bonds is considered the most significant step in detoxifica-

tion (Fig. 1). Both co-metabolic and bio-mineralization oforganophosphorus compounds by isolated bacteria have

been reported. A list of micro-organisms capable of degrad-

ing these compounds is presented in Table 2.

Hydrolysis of organophosphorus compounds leads to a

reduction in their mammalian toxicity by several orders of

magnitude. Since most of the research has been directed

towards detoxification, studies on the further metabolism of

the phosphorus containing products have not been exten-

sive. Hypothetical phospho-ester hydrolysis steps can be

postulated, yielding mono-ester and finally inorganic phos-

phate, but this pathway has not been specifically studied.

Analogous phospho-monoesterase and diesterase, which

degrade methyl and dimethyl phosphate, respectively, have

been reported inKlebsiella aerogenes(Wolfenden & Spence,

1967) and are produced only in the absence of inorganic

phosphate from the growth medium. The final enzyme in

the postulated degradative pathway is bacterial alkaline

phosphatase, which can hydrolyze simple monoalkyl phos-

phates and is also regulated by the level of phosphate

available to the cell (Wolfenden & Spence, 1967). A similar

mechanism of metabolism has been reported for phospho-

nates (Kerteszet al., 1994a). The way in which metabolism is

regulated depends very strongly on what role the organo-

phosphorus compound plays for the particular organisms

studied. Most often these compounds are used to supplyonly a single element (carbon, phosphorus or sulfur) and

the relevant gene cannot be expressed as a response to

starvation for another of these elements (Kertesz et al.,

1994a). For example, a strain of Pseudomonas stutzeri

isolated to utilize parathion as a carbon source released the

diethylphosphorothioanate products quantitatively and

could not metabolize them further, even when alternative

source of phosphorus or sulfur were removed (Daughton &

Hsieh, 1977). Similarly, a variety of isolates that could use

phosphorothionate and phosphorodithionate pesticides as a

sole source of phosphorus were unable to utilize these

compounds as a source of carbon (Rosenberg & Alexander,1979). Shelton (1988) isolated a consortium that could use

diethylthiophosphoric acid as a carbon source but was

unable to utilize it as a source of phosphorus or sulfur.

Kerteszet al. (1994a) explained possible underlying reasons

for this phenomenon. They suggested that the conditions

under which environmental isolates enriched were crucial in

selecting for strains not only with the desired degradative

enzyme systems but also with specific regulation mechan-

isms for the degradation pathways.

Table 1. History, toxicity and half-life of some organophosphorus

pesticides

Name Type

Year of

introduction

Mammalian

LD50

(mgkg1)

Half-life

soil

(days)

Chlorpyrifos Insecticide 1965 135163 10120

Parathion Insecticide 1947 210 30180

Methyl parathion Insecticide 1 949 330 25130

Glyphosate Herbicide 1971 35305600 30174

Coumaphos Acaricide 1952 1641 241400

Fenamiphos Nematicide 1967 610 2890

Monocrotophos Insecticide 1965 1820 4060

Dicrotophos Insecticide 1965 1522 4560

Diazinon Insecticide 1953 80300 1121

Dimethoate Insecticide 1955 160387 241

Fenitrothion Insecticide 1959 1700 1228

Ethoprophos Nematicide 1966 146170 330




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Table 2. Microorganisms isolated for the degradation of organophosphorus compounds

Compound Microorganisms Mode of degradation Reference

Chlorpyrifos Bacteria

Enterobactersp. Catabolic (C, P) Singhet al. (2003c)

Flavobacterium sp. ATCC27551 Co-metabolic Mallicket al. (1999)

Pseudomonas diminuta Co-metabolic Serdaret al. (1982)

Micrococcussp. Co-metabolic Guhaet al. (1997)Fungi

Phanerochaete chrysosporium Catabolic (C) Bumpuset al. (1993)

Hypholama fascicularae ND Bendinget al. (2002)

Coriolus versicolor ND Bendinget al. (2002)

Aspergillussp. Catabolic (P) Obojskaet al. (2002)

Trichoderma harzianum Catabolic (P) Omar (1998)

Pencillium brevicompactum Catabolic (P) Omar (1998)

Parathion Bacteria

Flavobacterium sp. ATCC27551 Co-metabolic Sethunathan & Yoshida (1973)


Pseudomonas stutzeri Co-metabolic Daughton & Hsieh (1977)

Arthrobacterspp. Co-metabolic Nelsonet al. (1982)

Agrobacterium radi obacter Co-metabolic Horneet al. (2002b)

Bacillusspp. Co-metabolic Nelsonet al. (1982)Pseudomonassp. Catabolic (C, N) Siddaramappaet al. (1973)

Pseudomonasspp. Catabolic (P) Rosenberg & Alexander (1979)

Arthrobactersp. Catabolic (C) Nelsonet al. (1982)

Xanthomonassp. Catabolic (C) Rosenberg & Alexander (1979)

Methyl parathion

Pseudomonassp. Co-metabolic Chaudryet al. (1988)

Bacillussp. Co-metabolic Sharmilaet al. (1989)

PlesimonasspM6 Co-metabolic Zhongliet al. (2001)

Pseudomonas putida Catabolic (C) Rani & Lalitha-kumari (1994)

Pseudomonassp. A3 Catabolic (C, N) Zhongliet al. (2002)

Pseudomonassp. WBC Catabolic (C, N) Yaliet al. (2002)

Flavobacterium balustinum Catabolic (C) Somara & Siddavattam (1995)

Glyphosate Bacteria

Pseudomonasssp. Catabolic (P) Kerteszet al. (1994a)

Alcaligene sp. Catabolic (P) Tolbotet al. (1984)

Bacillus megaterium2BLW Catabolic (P) Quinnet al. (1989)

Rhizobium sp. Catabolic (P) Liuet al. (1991)

Agrobacteriumsp. Catabolic (P) Wacketet al. (1987)

Arthrobactersp. GLP Catabolic (P) Pipkeet al. (1987)

Arthrobacter atrocyaneu s Catabolic (P) Pike & Amrhein (1988)

Geobacillus caldoxylosilyticusT20 Catabolic (P) Obojskaet al. (2002)

Flavobacterium sp. Catabolic (P) Balthazor & Hallas (1986)

Fungi

Penicillium citrium Co-metabolic Pothuluriet al. (1998)

Pencillium natatum catabolic (P) Pothuluriet al. (1992)

Penicillium chrysogenum Catabolic (N) Klimeket al. (2001)

Trichoderma viridae Catabolic (P) Zboinskaet al. (1992b)

Scopulariopsis spand Catabolic (P) Zboinskaet al. (1992b)

Aspergillus niger Catabolic (P) Zboinskaet al. (1992b)

Alternaria alternata Catabolic (N) Lipoket al. (2003)

Coumaphos

Nocardiodes simplexNRRL B24074 Co-metabolic Mulbry (2000)

Agrobacterium radi obacterP230 Co-metabolic Horneet al. (2002b)

Pseudomonas monteilli Co-metabolic Horneet al. (2002c)

Flavobacterium sp. Co-metabolic Adhyaet al. (1981)


Nocardiastrain B-1 Catabolic (C) Mulbry (1992)




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Chlorpyrifos

Chlorpyrifos (O,O-diethyl O-(3,5,6-trichloro-2-pyridyl)

phosphorothioate) is one of the most widely used insecti-

cides effective against a broad spectrum of insect pests of

economically important crops. It is effective by contact,ingestion and vapour action but is not systemically active. It

is used for the control of mosquitoes (larvae and adults),

flies, various soil and many foliar crop pests and household

pests. It is also used for ectoparasite control on cattle and

sheep. It has low solubility in water (2 mg L1) but is readily

soluble in most organic solvents. It has a high soil sorption

co-efficient (Racke, 1993) and is stable under normal storage

conditions. Chlorpyrifos is defined as a moderately toxic

compound having acute oral LD50; 135163 mg kg1 for rat

and 500 mg kg1 for guinea pig.

The environmental fate of chlorpyrifos has been studied

extensively. Degradation in soil involves both chemicalhydrolysis and microbial activity. The half-life of chlorpyr-

ifos in soil varies from 10 to 120 days (Getzin, 1981; Racke

et al., 1988) with 3,5,6-trichloro-2-pyridinol (TCP) as the

major degradation product. This large variation in half-life

has been attributed to different environmental factors, the

most important of which are soil pH, temperature, moisture

content, organic carbon content and pesticide formulation

(Getzin, 1981a, b; Chapman & Chapman, 1986). Initially,

the high rate of chlorpyrifos degradation in soils with

alkaline pH was attributed to chemical hydrolysis. Later,

Rackeet al. (1996) concluded that the relationship between

high soil pH and chemical hydrolysis was weak and that

other factors like soil silt content might be important in

determining environmental fate.

Unlike other organophosphorus compounds, chlorpyri-fos has been reported to be resistant to the phenomenon of

enhanced degradation (Rackeet al., 1990). There have been

no reports of enhanced degradation of chlorpyrifos since its

first use in 1965 until recently. It was suggested that the

accumulation of TCP, which has anti-microbial properties,

acts as a buffer in the soil and prevents the proliferation of

chlorpyrifos degrading microorganisms (Rackeet al., 1990).

However, Robertson et al. (1998) suggested that chemical

hydrolysis of chlorpyrifos and enhanced degradation of TCP

can result in loss of efficacy of the insecticide against

termites in sugar cane fields in Australia. Attempts to

introduce enhanced degradation in the laboratory or in thefield by repeated application have failed (Rackeet al., 1990;

Mallicket al., 1999).

In recent experiments, we found that the degradation of

chlorpyrifos was very slow in acidic soils but that the rate of

degradation increased considerably with an increase in soil

pH. However, in 90 days of incubation, there was no

difference between soils in release of 14CO2 from the

pyridine ring despite the large differences in degradation

rate. Repeated applications of chlorpyrifos did not affect

Monocrotophos

Pseudomonas spp. Catabolic (C) Bhadbhadeet al. (2002b)

Bacillus spp. Catabolic (C) Rangaswamy & Venkateswaralu (1992)

Arthrobacterspp. Catabolic (C) Bhadbhadeet al. (2002b)

Pseudomonas mendocina Catabolic (C) Bhadbhadeet al. (2002a)

Bacillus megaterium Catabolic (C) Bhadbhadeet al. (2002b)

Arthrobacter atrocyan eus Catabolic (C) Bhadbhadeet al. (2002b)Pseudomonas aeruginosaF10B Catabolic (P) Singh & Singh (2003)

Clavibacter michiganenseSBL11 Catabolic (P) Singh & Singh (2003)

Fenitrothion

Flavobacterium sp. Co-metabolic Adhyaet al. (1981)

Arthrobacter aurescen esTW17 Catabolic (C) Ohshiroet al. (1996)

Burkholderia sp. NF100 Catabolic (C) Hayatsuet al. (2000)

Diazinon

Flavobacterium sp. Catabolic (P) Sethunathan & Yoshida (1973)

Pseudomonas spp. Co-metabolic Rosenberg & Alexander (1979)

Arthrobacterspp. Co-metabolic Bariket al. (1979)

Chemical warfare agents

G Agent Pseudomonas diminuta Co-metabolic Mulbry & Rainina (1998)

Altermonasspp. Co-metabolic DeFranket al. (1993)

V Agent Pseudomonas diminuta Co-metabolic Mulbry & Rainina (1998)

Pleurotus ostreatus(fungus) Co-metabolic Yanget al. (1990)

Symbol in brackets after mode of degradation represents the type of nutrient that the pesticide provides to degrading microorganisms. C, carbon; N,

nitrogen; P, phosphorus.




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either the degradation rate or degradation kinetics, suggest-

ing that repeated treatment did not result in enhanced

degradation. Fumigation of soil samples completely inhib-

ited hydrolysis of chlorpyrifos, suggesting an involvement of

soil micro-organisms (Singhet al., 2003c). Chlorpyrifos has

been reported previously to be resistant to enhanced degra-

dation. Given the tremendous adaptability of the soilmicrobial community for degradation of a wide variety of

synthetic compounds, Racke et al. (1990) cited three possi-

ble reasons why a specific pesticide might not be susceptible

to enhanced degradation. One possibility is an inability of

the microflora to initiate degradation of the parent pesticide

easily. This may be due to factors such as steric hindrance of

enzymes by functional groups, electronic stability against

hydrolysis or lack of weak links in the molecule (Alexander,

1965; Niemiet al., 1987). The pesticide may also be unavail-

able for uptake and degradation by soil microorganisms due

to strong sorption to organic surfaces in the soil (Orgam

et al., 1985). However, these reasons cannot explain the

present results because chlorpyrifos is rapidly hydrolyzed by

the soil bacterial community in alkaline soils. The second

possibility is that the soil environmental conditions may in

some way inhibit the development or expression of en-

hanced degradation. This also cannot explain the present

results because repeated treatment of the same soil samples

resulted in enhanced degradation of fenamiphos (Singh

et al., 2003b). A third possibility is that the soil micro-

organisms cannot beneficially catabolize pesticide metabo-

lites. In these circumstances co-metabolism may occur (e.g.

hydrolysis of parent pesticides), but the microbial metabo-

lism of the degradation products is not possible. This is the

case with such relatively recalcitrant pesticides as DDT andalachlor, which are converted to products that are them-

selves quite resistant to further metabolism (Tiedje &

Hagedorn, 1975). From our experiments we concluded that

in high pH soils, the microbial community transforms

chlorpyrifos co-metabolically into TCP. However, TCP con-

tains three chlorine atoms on the pyridinol ring. To break

this ring, chlorine atoms have to be removed (Feng et al.,

1997), and free chlorine has toxic effects on the micro-

organisms. Thus TCP metabolism may be toxic to micro-

organisms. Similar results were obtained by Price et al.

(2001) in a field where degradation of chlorpyrifos was

strongly related with soil pH but degradation was mediatedby soil micro-organisms. Later, Singh et al. (2003c) sug-

gested that chlorpyrifos is degraded by non-specific and

non-inducible enzyme systems produced in high pH soils.

This suggests that chlorpyrifos is co-metabolically hydro-

lyzed to TCP and that because the TCP has toxic effects,

normally enhanced degradation does not occur. Although

Shelton & Doherty (1997) in their model proposed a

significant role of bioavailability in degradation of xenobio-

tics, the toxic effect of TCP seems to be a realistic explana-

tion of its resistance to enhanced degradation because TCP

has high water solubility and therefore is bioavailable for the

degradation. However, repeated treatment with chlorpyrifos

over many years in an Australian soil resulted in develop-

ment of some opportunist microorganisms with the cap-

ability to use the toxic compound as has been reported with

organochlorine compounds (Robertson et al., 1998; Singhet al., 2000). This adaptation can provide them with a

competitive advantage over other microbes in terms of

sources of energy. Further studies found higher copy num-

bers ofopd(organophosphate degrading) gene in higher pH

soils (Singhet al., 2003a, c).

In most cases described to date, the aerobic bacteria tend

to transform chlorpyrifos by hydrolysis to produce

diethylthiophosphoric acid (DETP) and TCP, which in turn

accumulate in the culture medium without further metabo-

lism. This transformation reaction removes chlorpyrifos and

its mammalian toxicity but yields compounds that are not

metabolized by the microorganisms that produce them

(Richins et al., 1997; Mallick et al., 1999; Horne et al.,

2002b; Wanget al., 2002b).

Chlorpyrifos has been reported to be degraded co-meta-

bolically in liquid media byFlavobacteriumsp. andPseudo-

monas diminuta, which were initially isolated from a

diazinon treated field and by parathion enrichment, respec-

tively (Sethunathan & Yoshida, 1973; Serdar et al., 1982).

However, these microbes do not utilize chlorpyrifos as a

source of carbon. A Micrococcus sp. was isolated from a

malathion enriched soil which was later reported to degrade

chlorpyrifos in liquid media (Guha et al., 1997). We have

isolated an Enterobacter sp. from a soil from Australia

showing enhanced degradation of chlorpyrifos. This bacter-ium degrades chlorpyrifos to DETP and TCP and utilizes

DETP as a source of carbon and phosphorus (Singh et al.,

2003c, 2004). Cook et al. (1978a) isolated several bacteria

from sewage sludge that were able to use dialkylthiopho-

sphonic acid as a sole source of phosphorus. One of these

organisms,Pseudomonas acidovorans, was able to use DETP

as a sole source of sulfur (Cook et al., 1980). Another

significant observation was the utilization of organopho-

sphorus insecticides as a source of phosphorus byEntero-

bactersp. (Singhet al., 2003c, 2004). Sethunathan & Yoshida

(1973) isolated aFlavobacteriumsp. that could use diazinon

as a source of carbon. However,Flavobacteriumwas not ableto use other organophosphorus pesticides as a source of

either phosphorus or carbon. Similarly, a variety of isolates

that could use phosphorothionate or phosphorodithionate

compounds as a sole source of phosphorus were unable to

degrade these compounds as a source of carbon (Rosenberg

& Alexander, 1979). Shelton (1988) isolated a consortium

that could use DETP as a carbon source but was unable to

degrade it when presented as source of phosphorus or sulfur.

It is believed that the conditions under which environmental




8/44

isolates are enriched are crucial in selecting for strains not

only with the desired degradative enzymes systems, but also

with the specific regulation mechanisms for the degradation

pathways (Kerteszet al., 1994a).

Studies on further metabolism and identification of

intermediate products of the phosphorus containing pro-

ducts have not been extensive. The postulated pathway stepsinclude hydrolysis, yielding monoester and finally inorganic

phosphate (Fig. 2). Bacterial phosphodiesterase has been

purified from a wide range of organisms including Escher-

ichia coli (Imamura et al., 1996), Haemophilus influenzae

(Macfadyen et al., 1988), and Burkholderia caryophylli

PG2982 (Dotsonet al., 1996). The phosphodiesterase from

the first two bacteria are similar in sequence and both

moderate intracellular cyclic AMP levels. However, the

phosphodiesterase from B. caryophilli has a different se-

quence from that in the first two bacteria (Dotson et al.,

1996). This enzyme could not be assigned a clear function

but was thought to play a role in xenobiotic degradation

pathways because it degraded glycerol glyphosate. However,

until recently no phosphodiesterase had been isolated or

characterized which could utilize xenobiotic degradation

products such as diethyl phosphate and diethyl phospho-

nate. A novel phosphodiesterase was isolated and cloned

from Delftia acidovorans which has both mono- and di-

esterase activity (Tehara & Keasling, 2003). This enzyme

allows D. acidovorans to use diethyl phosphonate as a sole

source of phosphorus under phosphorus limiting condi-

tions. The final enzyme in the postulated degradative path-

way is alkaline phosphatase, which can hydrolyze simple

monoalkyl phosphates (Neidhardtet al., 1996).

Since only one bacterium has been isolated so far whichcan degrade TCP in liquid medium, little literature is

available on microbial metabolism of TCP. Feng et al.

(1997) isolated a Pseudomonas sp. which can mineralize

TCP in liquid medium. Later the same group, on the basis of

combined experiments with photolysis and microbial de-

gradation, suggested that TCP was metabolized by a Pseu-

domonas sp. by a reductive dechlorination pathway (Feng

et al., 1998). In this pathway, TCP is first reductively

dechlorinated into chlorodihydro-2-pyridone, which is

further dechlorinated to tetra-hydro-2-pyridone. Ring clea-

vage of this compound resulted in formation of maleamide

semialdehyde, which is metabolized to water, carbon diox-ide, and ammonium ions. Microbial degradation of analo-

gous compounds such as pyridine and hydroxypyridine has

been researched and reviewed extensively (Shukla, 1984;

Sims & OLoughlin, 1989; Kaiseret al., 1996). Several micro-

organisms were reported to degrade hydroxypyridine (Kai-

ser et al., 1996). Cain et al. (1974) reported that 2- or 3-

hydroxypyridine was oxidized to 2,5-dihydroxypyridine and

production of maleamic acid occurred later through ring

cleavage. Oxygen atoms used to transform 4-hydroxypyr-

idine via 3,4-dihydroxypyridine were derived from water

molecules by hydroxypyridine hydrolase (Watson et al.,

1974). It is likely that TCP is metabolized in a similar

manner as one of the metabolites of TCP was identified to

have similar structure to 2-hydroxypyridine.

Fungal mineralization of chlorpyrifos byPhanerochaete

chrysosporium was reported by Bumpus et al. (1993).Chlorpyrifos was hydrolyzed and then the pyridinyl ring

underwent cleavage before being converted to carbon diox-

ide and water. Degradation of chlorpyrifos in biobed

composting substrate by two other white-rot fungi,Hypho-

loma fascicularae and Coriolus versicolor, was observed

(Bending et al., 2002). Degradation of a wide range of

xenobiotic compounds by white-rot fungi is well documen-

ted (Kuhadet al., 1997; Singh & Kuhad, 1999, 2000; Singh

et al., 1999). These organisms have been reported to degrade

several persistent aromatic compounds by ring cleavage

(Armenante et al., 1994; Reddy & Gold, 2000). The multi-

step pathway of pentachlorphenol degradation by the white-

rot fungusPhanerochaete chrysosporiumis initiated by lignin

peroxidase and manganese peroxidase, producing tetra-

chloro-1-4-benzoquinone, which is further metabolized by

two parallel but cross-linked pathways. The tetrachloroben-

zoquinone is reduced to tetrachlorodihydroxybenzene,

which can undergo four successive dechlorinations to pro-

duce 1,4-hydroquinone. This is then hydroxylated to pro-

duce the final aromatic metabolite, 1,2,4-trihydroxybenzene.

Alternatively the tetrachlorobenzoquinone converts to

2,3,5-trichlorotrihydroxybenzene, which undergoes succes-

sive reductive dechlorination to produce 1,2,4-trihydroxy-

benzene. At several points, hydroxylation reaction converts

chlorinated dihydroxybenzene to chlorinated trihydroxy-benzene, linking two pathways. The 1,2,4-trihydroxyben-

zene is ring cleaved to produce CO2 and water (Reddy &

Gold, 2000). Mineralization of TCP by white-rot fungi is

possible via reductive de-chlorination. White-rot fungi have

been reported previously to use this transformation step to

degrade other chlorinated compounds such as pentachlor-

ophenol (Aiken & Logan, 1996) and hexachlorocyclohexane

(Mouginet al., 1996; Singh & Kuhad, 1999, 2000). Degrada-

tion of several polychlorinated compounds by white-rot

fungi suggests that they produce a range of isoenzymes with

a wide range of substrate specificity. Several species of

Aspergillus, Trichoderma harzianum and Penicillium brevi-compactumwere reported to utilize chlorpyrifos as sources

of phosphorus and sulfur (Omar, 1998) (Table 1). On the

basis of the above discussion, the authors propose possible

pathways for microbial degradation of chlorpyrifos (Fig. 2).

Parathion

Parathion (O,O-diethyl-O-p-nitrophenyl phosphorothio-

ate) is one of the most toxic insecticides registered with the




9/44

US Environmental Protection Agency (EPA). Extreme toxi-

city with ease of exposure has resulted in numerous human

and non-target species deaths in several developing coun-

tries (McConnellet al., 1999). The microbial degradation of

parathion has received extensive attention among the orga-

nophosphorus compounds because of its widespread use

and the ready detection of its hydrolytic product (p-nitro-

phenol). Parathion is rapidly degraded in biologically active

Fig. 2. Proposed pathways for chlorpyrifos

degradation by microorganisms. The scheme is

based on articles cited in the text. When the

conversion of one compound to another is

believed to occur through a series of intermediates, the steps are indicated by dotted

arrows. DETP, diethylthiophosphate; TCP,

trichloropyridinol.




10/44

soil. A proportional increase in the bacterial population in

soils was observed with an increase in the concentration of

parathion added (Nelson, 1982). Flooded soil conditions

favoured hydrolysis of parathion and release of 14CO2from

ring labelled parathion in the rhizosphere of rice seedlings

(Reddy & Sethunathan, 1983).

Several species of bacteria have been isolated either fromparathion enrichment or other organophosphate enriched

environments, which can hydrolyze parathion (Table 2)

(Munneckeet al., 1982; Kertesz et al., 1994a; Racke et al.,

1996). Both mineralization, where parathion was used as a

source of carbon (Munnecke & Hsieh, 1976; Rani & Lalitha-

kumari, 1994) or phosphorus (Rosenberg & Alexander,

1979), and co-metabolic hydrolysis (Serdar et al., 1982;

Horne et al., 2002b) have been reported. Sethunathan &

Yoshida (1973) isolated the first organophosphorus degrad-

ing bacterium,Flavobacteriumsp., that could degrade para-

thion and diazinon. Siddaramappa et al. (1973) isolated a

Pseudomonas sp. that was able to hydrolyze parathion and

utilize the hydrolysis product p-nitrophenol as a carbon or

nitrogen source. Later, P. stutzeri was isolated, which can

hydrolyze parathion although p-nitrophenol was metabo-

lized by a separate bacterium (Daughton & Hsieh, 1977).

Rosenberg & Alexander (1979) isolated two Pseudomonas

ssp. that were able to hydrolyze a number of organopho-

sphorus compounds including parathion, and to use the

ionic cleavage products as a sole source of phosphorus.

Several species of Bacillus and Arthrobacter have been

isolated that were capable of hydrolyzing parathion; one of

theArthrobacterstrains was also able to utilizep-nitrophenol

as a sole source of carbon (Nelson, 1982). A Pseudomonassp.

and a Xanthomonas sp. were isolated which can hydrolyzeparathion and can further metabolizep-nitrophenol (Tche-

letet al., 1993). AMoraxellasp. can usep-nitrophenol as the

sole source of carbon and nitrogen (Spain & Gibson, 1991).

This bacterium degrades p-nitrophenol to p-benzoquinone

using the enzymep-nitrophenol monooxygenase. p-Benzo-

quinone is transformed to hydroquinone by a reductase

(Spain & Gibson, 1991). Candida parapsilosis has been

reported to produce hydroquinine 1,2-dioxygenase, which

converts hydroquinone to cis,trans-4-hydroxymuconic

semialdehyde. This is then metabolized to maleylacetate by

semialdehyde dehydrogenase. Maleylacetate is converted to

3-oxoadipate by a reductase, which is finally metabolized tointermediary metabolites of the tricarboxylic acid (TCA)

cycle (Carnett, 2002). A Pseudomonas putida strain was

found to metabolize p-nitrophenol to hydroquinone and

1,2,4-benzenetriol, which was further cleaved by benzene-

triol oxygenase to maleylacetate (Rani & Lalitha-kumari,

1994). A similar pathway ofp-nitrophenol degradation was

reported in Pseudomonas cepacia that can utilize p-nitro-

phenol as a source of carbon and nitrogen (Prakash et al.,

1996).

A different pathway of degradation was reported in

Arthrobactersp. strain JS443 and Arthrobacter protophormiae

RHJ100 wherep-nitrophenol was mineralized viap-nitroca-

techol. Nitrocatechol is converted to 1,2,4-benzenetriol by

benzotriol dehydrogenase, which in turn is directly con-

verted to maleylacetate by benzotriol dioxygenase (Jainet al.,

1994; Bhushanet al., 2000a; Chauhanet al., 2000). Recently,a consortium of two Pseudomonas ssp. (strains S1 and S2)

was isolated which can also metabolize p-nitrophenol via

p-nitrocatechol (Qureshi & Purohit, 2002). The analogous

compound 3-methyl-4-nitrophenol has also been reported

to be metabolized by Ralstonia sp. via catechol formation

(Bhushan et al., 2000b). A Nocardia sp. was reported to

producep-nitrophenol-2-hydroxylase, which catalyzes trans-

formation of p-nitrophenol to p-nitrocatechol (Mitra &

Vaidyanathan, 1984). A mono-oxygenase from a Moraxella

sp. that releases nitrite fromp-nitrophenol has been partially

purified (Spain & Gibson, 1991). A soluble nitrophenol

oxygenase was purified from P. putida B2 that converts

ortho-nitrophenol to catechol and nitrite (Zeyer & Kocher,

1988). A novel monooxygenase was characterized from

Bacillus sphaericus that catalyzes the first two steps of the

degradation ofp-nitrophenol viap-nitrocatechol and benzo-

triol. This enzyme consists of two components, a reductase

and oxygenase, and catalyzes two sequential mono-oxygena-

tion reactions that convertp-nitrophenol to benzotriol. The

first reaction converts p-nitrophenol to p-nitrocatechol and

the second removes the nitro group (Kadiyala & Spain,

1998). A pentachlorophenol degrading Sphingomonas

sp. UG30 was found to degradep-nitrophenol. A pentachloro-

phenol-monooxygenase was purified from this bacterium

that can catalyze the hydroxylation of p-nitrocatechol tobenzotriol (Leung et al., 1999). A hydroxyquinol (benzo-

triol) ring cleavage dioxygenase was isolated and character-

ized from p-nitrophenol degrading Arthrobacter sp. strain

JS443. The gene encoding this dioxygenase (npdB) was

found to be in the same gene cluster as reductase ( npdA1)

and oxygenase (npdA2) components of the p-nitrophenol

mono-oxygenase, maleylacetate reductase (npdC), and a

regulatory protein (npdR) (Zylstra et al., 2000; Parales

et al., 2002). Rhodococcus strain PN1 and Rhodococcus

erythropolis HL PM-1, which degrade 2,4-dinitrophenol

and p-nitrophenol, were reported to contain an npdgene

cluster including npdC (encoding hydride transferase I),npdG (encoding the NADPH-dependent F420 reductase)

andnpdI (encoding hydride transferase II). It was observed

that npdG and npdI genes have the same function as the

homologous genes (Heisset al., 2003). Recently, a novel gene

called orf243 was reported from Flavobacterium sp. orf243

which is transposon based and is linked with the opdgene

(Siddavattamet al., 2003). This gene encodes a protein with

homology to a family of aromatic compound hydrolases and

is able to degradep-nitrophenol.




11/44

Although in most of the studies on microbial degradation

of parathion, the first reaction was hydrolysis of the phos-

photriester bond, there have been reports of different

degradation pathways. In one study, degradation of para-

thion by a mixed culture and a Bacillussp. (Sharmila et al.,

1989) was shown to occur by reduction of the nitrogroup

that was later hydrolyzed top-aminophenol. Another reportof conversion of parathion to paraoxon before hydrolysis of

phosphotriester bond was reported in a mixed bacterial

culture (Tomlin, 2000).

Studies on the degradation of methyl parathion (O,O-

dimethyl-O-p-nitrophenyl phosphorothioate) have also

been reported. Methyl and ethyl parathion have identical

chemical structures except for the ethyl groups of the P

chain of parathion, which are replaced by methyl groups as

evident by the name of the compound. A Pseudomonas sp.

was isolated that can co-metabolically degrade methyl para-

thion (Chaudryet al., 1988). Rani & Lalitha-kumari (1994)

isolatedP. putidathat could hydrolyze methyl parathion and

utilizep-nitrophenol as a source of energy. ABacillussp. was

reported to degrade methyl parathion by both hydrolysis

and nitro group reduction (Sharmila et al., 1989). Utiliza-

tion of methyl parathion byFlavobacterium balustinum as

the sole source of carbon was observed earlier (Somara &

Siddavattam, 1995). In this bacterium the opd gene was

found to be linked with a novel gene involved in degradation

ofp-nitrophenol (Siddavattamet al., 2003). Degradation of

methyl parathion by a Pseudomonas sp. in soil and on

sodium alginate beads was reported (Ramanathan & La-

lithakumari, 1996). Co-metabolic degradation of methyl

parathion byPlesimonassp. strain M6 was observed (Zhon-

gli et al., 2001) which was mediated by a novel degradinggene. They also isolated Pseudomonas sp. A3 which can

utilizep-nitrophenol as sole source of carbon and nitrogen.

This isolate can also utilize a series of aromatic compounds

as a sole source of carbon (Zhongli et al., 2002). Another

strain ofPseudomonas sp. WBC was isolated from polluted

soils around a Chinese pesticide factory. The isolate was

capable of complete degradation of methyl parathion and

could utilize it as sole source of carbon and nitrogen (Yali

et al., 2002). The hydrolysis product of methyl parathion is

also p-nitrophenol, for which the degradation pathways

have already been described. The different proposed path-

ways of parathion and methyl degradation are presentedin Fig. 3.

Glyphosate

Glyphosate (N-(phosphonomethyl) glycine) is a globally

used broad-spectrum herbicide. It is a representative of the

phosphonic acid group of compounds, which is character-

ized by a direct carbon to phosphorus (CP) bond. The CP

linkage is chemically and thermally very stable and renders

the molecule much more resistant to non-biological degra-

dation in the environment than its analogues with O-P

linkage (Hayes et al., 2000). Mode of action of glyphosate

includes inhibition of the plant enzyme 5-enol-pyruvyl-

shikimate-3-phosphate synthase, which catalyzes synthesis

of aromatic amino acids (Fisher et al., 1984; Cole, 1985).

Glyphosate is moderately persistent with a half-life of30170 days (Tomlin, 2000). Microbial degradation is

considered to be the most important of the transformation

processes controlling its persistence in soil (Araujo et al.,

2003). It was observed that mineralization of glyphosate is

related to both the activity and biomass of soil micro-

organisms (Wiren-Lehret al., 1997). Microbial degradation

of glyphosate produces the major metabolite aminomethyl

phosphonic acid and ultimately leads to the production of

CO2, phosphate and water (Forlani et al., 1999; Araujo

et al., 2003). Several species of bacteria have been isolated

from previously treated and untreated environments, which

can degrade glyphosate either co-metabolically or as a

source of phosphorus. There has been no report of the

utilization of glyphosate as a source of carbon or nitrogen

(Dick & Quinn, 1995). Several species ofPseudomonashave

been isolated which can degrade glyphosate (Moore et al.,

1983; Tolbot et al., 1984; Jacob et al., 1988; Quinn et al.,

1989). Similarly, a Flavobacterium sp. (Balthazor & Hallas,

1986), an Alcaligenes sp. (Tolbot et al., 1984), Bacillus

megateriumstrain 2BLW (Quinnet al., 1989), several species

of Rhizobium(Liuet al., 1991), three species of Agrobacter-

ium (Wacket et al., 1987; Liu et al., 1991) and an Arthro-

bacter sp. (Pipke et al., 1987) have also been reported to

degrade this herbicide (Table 2).

Three different pathways for CP bond cleavage havebeen reported for the use of phosphonate as a source of

phosphorus for growth.

The phosphonatase pathway is involved in degra-

dation of alpha carbon substituted phosphonates, which are

primarily naturally occurring phosphonates such as

2-aminoethylphosphonates that have been reported in

Bacillus cereus (Lee et al., 1992b), and Pseudomonas aeru-

ginosa (Lacoste et al., 1993), Salmonella typhimurium and

several other organisms (Jiang et al., 1995). In a two-step

process, this pathway leads to the cleavage of the CP bond

by a hydrolysis reaction requiring an adjacent carbonyl

group. 2-Aminoethylphosphonate is converted to phosp-honoacetaldehyde by a specific transaminase, which is

further degraded to acetaldehyde by phosphonatase.

The CP lyase pathway is involved in the cleavage of

both substituted and unsubstituted phosphonates such as

methylphosphonates (Lee et al., 1992b).The phospho-

noacetate hydrolase pathway specifically degrades phospho-

noacetate and appears to have evolved for phosphonate use

as a carbon source. This enzyme catalyzes the hydrolysis of

phosphonoacetate ;to acetate and inorganic phosphonates




12/44

via metal cation-assisted PC bond cleavage (McMullan &

Quinn, 1994; McGrath et al., 1995). Glyphosate has beenfound to be degraded by the second of these pathways.

Two different pathways of glyphosate degradation are

presented in Fig. 4.Arthrobactersp. GLP-1 andPseudomonas

sp. PG2982 degrade glyphosate by initial cleavage of the CP

bond, resulting in the production of sarcosine (N-methyl-

glycine) by CP lyase activity (Moore et al., 1983; Shinabar-

ger & Braymer, 1984; Pipkeet al., 1987; Liuet al., 1991; Dick

& Quinn, 1995). Rhizobium meliloti has also been reported

to degrade glyphosate by this pathway but, unlike other

bacteria, it has only one CP lyase, which is able to degrade a

wide range of phosphonates (Park & Hausinger, 1995). Thesarcosine formed is further degraded to the amino acid

glycine and a C1-unit, which is incorporated into purines,

and the amino acids serine, cysteine, methionine and

histidine (Pipke et al., 1987). The second pathway involves

the conversion of glyphosate to aminomethylphosphonic

acid (AMPA) by the loss of a C2unit. This compound is then

dephosphorylated by CP lyase and further broken down by

subsequent steps to methylamine and formaldehyde (Pike &

Amrhein, 1988; Lerbset al., 1990). An identical pathway has

Fig. 3. Different pathways of parathion and

methyl parathion degradation by microorgan-

isms. When the conversion of one compound to

another is believed to occur through a series of

intermediates, the steps are indicated by dotted

arrows. DATP, dialkylthiophosphate.




13/44

been observed inArthrobacter atrocyaneus(Pike & Amrhein,

1988) and Flavobacterium sp. (Balthazor & Hallas, 1986;

Pipke et al., 1987). Recently, a thermophile, Geobacilluscaldoxylosilyticus T20 was isolated from a central heating

system which also degrades glyphosate by this pathway,

utilizing the compound as a sole source of phosphorus

(Obojskaet al., 2002). A halophilic bacterium,Chromohalo-

bacter marismortui, isolated from soil beneath a road

gritting salt pile was capable of utilizing several organopho-

sphonates including aminomethyl phosphonic acid as a

source of phosphorus (Hayes et al., 2000). Utilization of

aminoalkylphosphonates as a source of nitrogen by different

bacterial isolates has been reported (McMullan & Quinn,

1994; Ternana & McMullan, 2000). Pseudomonas fluorescens

was reported to utilize a diverse range of organophospho-nates as sources of carbon, nitrogen and phosphorus

(Zboinska et al., 1992a). A strain ofKluyveromyces fragilis

has been shown to utilize AMPA as a source of nitrogen

(Ternana & McMullan, 2000). Strains ofStreptomyceswere

also reported to degrade and utilize several organopho-

sphonate compounds as sources of carbon and nitrogen.

These strains were capable of degrading glyphosate in

phosphate-free media via CP bond cleavage accompanied

by sarcosine formation (Obojska et al., 1999).Streptomyces

morookaensis DSM 40565 could degrade aminoalkylpho-

sphonate as a sole source of nitrogen and phosphorus

(Obojska & Lejczak, 2003). Alkyl amines are intermediatedegradation products for several xenobiotics such as carbo-

furan, atrazine, and monocrotophos and have been reported

to serve as a source of energy for different micro-organisms

(Strong et al., 2002). Use of methylamine as a source of

carbon is widespread in nature (Hanson & Hanson, 1996;

Trabueet al., 2001).

Fungi play an important role in degradation of xenobio-

tics and biospheres (Pothuluri et al., 1998, 1992) including

glyphosate. Probably the first fungal degradation of glypho-

sate byPenicillium citrinumwas reported by Zboinskaet al.

(1992b).Penicillium notatum can utilize the herbicide as a

source of phosphorus and can degrade it by the amino-methyl phosphonic acid pathway (Bujacz et al., 1995).

Strains ofTrichoderma harzianum, Scopulariopsis spandand

Aspergillus niger were able to degrade glyphosate and

aminomethyl phosphonic acid in the laboratory (Krzysko-

Lupicka et al., 1997). The first report of utilization of

glyphosate as a source of nitrogen by a microorganism was

reported forPenicillium chrysogenum (Klimeket al., 2001).

The fungal cells were found to lack detectable nitrogen

reductase activity and therefore this isolate seemed to lack

Fig. 4. Pathways of microbial degradation for

glyphosate. AMPA, aminomethyl phosphonic

acid.




14/44

the ability to convert nitrate to ammonium. Recently,

Alternaria alternata, a plant pathogen, was found to utilize

glyphosate as a source of nitrogen (Lipoket al., 2003).

The above observations suggest that glyphosate is de-

graded by several soil microorganisms, and different steps of

the degradation involve different microorganisms which

utilize different degradation products as different sources ofenergy. The possible pathways of glyphosate degradation are

presented in Fig. 4.

Coumaphos

Coumaphos (O,O-diethyl-O-(3-chloro-4-methyl-2-oxo-

2H-1-benzo-pyran-7-yl) phosphorothioate) is used as an

acaricide for the control of cattle ticks. It is widely used by

different government agencies for tick eradication and

quarantine purposes. The primary tool used in the eradica-

tion programme is a series of dipping vats placed at border

crossing points. The cattle are induced to jump into the deep

end of the vat, resulting in their complete immersion in

coumaphos. They then swim the length of the vat and climb

out to other end. There are around 42 vats in the USA alone

and each vat contains about 15 000 L of coumaphos suspen-

sion at the rate of 1600 mg L1 (42% active ingredient, a.i.)

(Shelton & Somich, 1988; Mulbryet al., 1998). The vats are

cleaned and recharged every 2 years to keep the concentra-

tion of acaricide at a desirable level. These operations

generate approximately 460 000 L of concentrated insecti-

cide waste yearly in USA alone (Mulbry et al., 1996). A

similar programme within Mexico is thought to produce a

much larger volume. Coumaphos is comparatively persis-

tent in soil, with a half-life of about 300 days (Kearneyet al.,1986) and it possesses a very high mammalian toxicity.

Because of these characteristics, it requires a safe and

effective method for disposal. Rapid degradation of couma-

phos was observed in several cattle-dipping vats, resulting in

loss of efficacy against cattle ticks (Shelton & Karns, 1988).

Under aerobic conditions, experiments with radiolabelled

coumaphos demonstrated that the aromatic portion of the

molecule is susceptible to mineralization by bacteria in

problematic vat dips (loss of efficacy). Three morphologi-

cally distinct bacteria (designated B-1, B-2 and B-3) that

could metabolize coumaphos were isolated from a problem

vat dip (Shelton & Somich, 1988). All these bacteria hydro-lyzed coumaphos to DETP and chlorferon. Chlorferon was

further metabolized by B-1 and B-2 to a-chloro-b-methyl-

2,3,4-trihydroxy-trans-cinnamic acid (CMTC). Further ex-

periments demonstrated that B-1 was capable of mineraliz-

ing and incorporating the aromatic portion of the

coumaphos molecule into biomass, but this was inhibited

by the accumulation of metabolites that was due apparently

to the inefficient metabolism of a chlorinated intermediate.

Combination of B-1 with another organism from the vat,

designated strain B-4, which metabolized these inhibitory

products, yielded a stable two-member consortium able to

grow at the expense of coumaphos (Shelton & Haperman-

Somich, 1991). No further study on the degradation path-

way or metabolite identification has been carried out.

Ralstonia sp. LD35 has been reported to degrade an

analogous compound, 3,4-dihydroxycinnamic acid via ben-zoic acid (Gioia et al., 2001). A similar breakdown pathway

for the propenoic side chain of substituted cinnamic acid

molecule,p-coumaric acid, has been observed inPseudomo-

nassp. (Tse et al., 2004) and Acinetobacter strains (Delneri

et al., 1995). These bacteria use p-coumaric acid as the

source of carbon. In the first step, they convert p-coumaric

acid intop-hydroxybenzoic acid which is then transformed

to protocatechuic acid and integrated to the TCA cycle via

the b-ketodipate pathway. Many bacteria degrade substi-

tuted cinnamic acid by decarboxylation of side chains.

Enzymes and genes responsible for such degradation have

been purified and characterized (Degrassi et al., 1995;

Barthelmebs et al., 2000). Streptomyces setonii (Sutherland

et al., 1983) and Rhodopseudomonas palustris (Harwood &

Gibson, 1988) have been shown to degrade cinnamic and 4-

coumaric acids to their corresponding benzoic acid deriva-

tives. Several other bacteria follow the same pathway for

degradation of substituted cinnamic acids. Monooxygenase

and dioxygenase catalyze the formation of the 2-, 3-, and 4-

hydroxy derivatives as substituted acid and/or substituted

catechol (Penget al., 2003).

The b-oxidation pathway has been proposed for the

degradation of substituted cinnamic acids byPseudomonas

putida(Zenket al., 1980). This pathway, which is analogous

to the b-oxidation of fatty acids, is thought to includethiolytic cleavage of 4-hydroxy-3-methoxy-b-ketopropinyl-

CoA to yield acetyl CoA and vanillyl CoA, which is catalyzed

byb-ketoacyl CoA thiolase. The pathway subsequently leads

to ring fission and requires several co-factors including ATP,

CoA and NAD1 (Zenk et al., 1980). Under anaerobic

conditions, coumaphos undergoes reductive dechlorination

to form potasan (Mulbryet al., 1998).

Nocardia sp. strain B-1 was reported to degrade couma-

phos by a different gene enzyme system to the known opd

gene (Mulbry, 1992). Another microorganism, Nocardiodes

simplex NRRL B-24074, was found to have a distinct

enzymes system for coumaphos degradation (Mulbry,2000). Horne et al. (2002b) isolated an Agrobacterium

radiobacter P230 capable of hydrolyzing coumaphos from

an enrichment culture containing organophosphorus as the

sole source of phosphorus. This bacterium degrades couma-

phos by hydrolysis of the phosphotriester bond.Pseudomo-

nas monteilli was isolated which can hydrolyze coumaphos

as well as its oxo analogue coroxon but it can utilize only

coroxon as a sole source of phosphorus, not coumaphos or

its hydrolysis product DETP. This bacterium degrades




15/44

coumaphos and diazinon but not parathion (Horne et al.,

2002a). Coumaphos is degraded by the other microorgan-

isms likeFlavobacteriumsp. (Sethunathan & Yoshida, 1973),

P. diminuta (Serdar et al., 1982), andEnterobacter sp. B-14

(Singhet al., 2004), which were isolated for their ability to

degrade other organophosphorus compounds. This obser-

vation suggests that these microorganisms produce severalisoenzymes or broad-specificity enzymes that can degrade a

range of organophosphorus compounds. The proposed

pathway of microbial degradation of coumaphos is shown

in Fig. 5.

Fenamiphos

Fenamiphos (ethyl 4-methylthio-m-tolyl isopropylpho-

sphoramidate) is an organophosphorate used extensively

for the control of soil nematodes. It is systemic, active

against ecto- and endo-parasitic, cyst forming and root-

knot nematodes, and is recommended for application at520kg a.i.ha-1. Its solubility at room temperature is

700mgL1 water. The acute oral LD50 is 15.319.4

mgkg1 for rats, 10mg kg1 for dogs and 75100 mg kg1

for guinea pigs (Tomlin, 2000).

Fig. 5. Proposed pathways for microbial

degradation of coumaphos. The scheme is based

on articles cited in the text. When the conver

sion of one compound to another is believed

to occur through a series of intermediates, the

steps are indicated by dotted arrows. DETP,

diethylthiophosphate; CMTC, chloromethyl

trihydroxy cinnamic acid.




16/44

Although, there have been reports of enhanced degrada-

tion of fenamiphos, the mechanism of degradation has

received little attention. Fenamiphos is oxidized rapidly to

fenamiphos sulfoxide (FSO) which in turn is oxidized to

fenamiphos sulfone (FSO2). As FSO and FSO2, have nema-

ticidal activity and toxicity similar to fenamiphos (Waggoner

& Khasawinah, 1974), degradation and persistence studiesusually include estimation of total toxic residue, which is the

combination of the two oxidation products along with

parent compound. The half-life in soil for fenamiphos and

its metabolites (total toxic residues) varies from 30 days to 90

days (Johnson, 1998). More rapid rates of degradation in soil

repeatedly treated with the fenamiphos in the laboratory

have been reported (Chung & Ou, 1996) and enhanced

degradation of fenamiphos in the field has been observed in

many countries (Stiriling et al., 1992; Smelt et al., 1996;

Meghrajet al., 1999). It was suggested that 34 years were

necessary before the accelerated degradation of fenamiphos

declined in a sandy soil in a temperate region (Ou, 1991).

Fenamiphos rapidly disappears from both enhanced and

non-enhanced soils but FSO2 is rarely formed in enhanced

soils (Ou, 1991). This suggests that enhanced bio-degrada-

tion of fenamiphos total toxic residue was due to an increase

in the disappearance rate of FSO in soil samples collected

from field sites treated one or two consecutive times with

fenamiphos (Davis et al., 1993). In a recent study of soil

samples from a field in the UK, which had similar physical

characteristics except for soil pH, the degradation rate of

fenamiphos increased with the increase in pH. Repeated

application of fenamiphos slowed down the rate of degrada-

tion in acidic soils, and in the neutral pH soil, three

consecutive treatments did not result in the development ofenhanced degradation of fenamiphos. However, in the two

alkaline soils, a second treatment with fenamiphos led to

enhanced degradation (Singh et al., 2003b). Chung & Ou

(1996) have tried to shed light on the mechanism of

fenamiphos degradation in soils that showed enhanced

degradation. They reported that fenamiphos is degraded

into FSO which in turn is rapidly degraded into FSO-

phenol, which is subsequently mineralized into CO2. There-

fore in enhanced soil, degradation of fenamiphos (total toxic

residue) is rapid because it misses one step, FSO to FSO2. In

enhanced UK soils, fenamiphos was rapidly oxidized to FSO,

which in turn, was quickly degraded. The major fenamiphosmetabolites identified were FSO and FSO-phenol. No FSO2was detected in the enhanced soil samples (Singh et al.,

2003b). However, in two Australian soils, a different me-

chanism of fenamiphos degradation was observed where the

nematicide was directly converted to fenamiphos phenol,

suggesting that the first oxidation step was replaced by

hydrolysis (Singhet al., 2003b).

Ou & Thomas (1994) isolated the first microbial con-

sortium with six different bacterial species that degraded

fenamiphos in liquid culture. A pure culture ofBrevibacter-

ium sp. MM1 was isolated which hydrolyzed fenamiphos

and its hydrolysis products but did not utilize these chemi-

cals as energy sources (Megharaj et al., 2003). Two different

consortia from Australian soils, made up of five and four

different bacterial strains, were isolated [B. K. Singh, un-

published]. Both consortia could utilize fenamiphos as solesources of carbon and nitrogen. In contrast to the con-

sortium isolated by Ou & Thomas (1994), the two Austra-

lian consortia (CRF and BEP) did not require any

supplementary nutrient source for fenamiphos degradation

and were active in liquid media in the absence of mineral

surfaces (Singhet al., 2003b). These microbial systems were

found to mineralize fenamiphos or its oxidative metabolites

by hydrolysis as a first step. The hydrolytic product fenami-

phos phenol, FSO-phenol or fenamiphos sulfone phenol

(FSO2-OH) can be further degraded by desulfonation. Three

modes of desulfonation are reported for aromatic sulfo-

nates: desulfonation (a) before, (b) during or (c) after ring

cleavage (Kerteszet al., 1994a). Mode (a) is considered to be

most common pathway of desulfonation in the environ-

ment. In this pathway, the target compound is oxygenated

by a multi-component oxygenase, yielding an unstable

sulfono cis-diol, which then spontaneously re-aromatizes

to the corresponding catechol with the loss of sulfite. An

enzyme which catalyzes this reaction in toluene sulfonate

and benzene sulfonate has been isolated from an Alcaligenes

sp. (Thurnheer et al., 1986, 1990). In Pseudomonas putida

S-313, a broad-spectrum monooxygenolytic sulfonatase

catalyzes the conversion of sulfonate to a phenol with

incorporation of one oxygen atom from molecular oxygen

(Kerteszet al., 1994b).Alcaligenessp. strain O-1 is reportedto contain two different desulfonative pathways where the

initial desulfonation is catalyzed by different dioxygenase

enzyme systems. One enzyme system can degrade 2-amino-

benzenesulfonate, benzene sulfonate and 4-toluene sulfo-

nate but the other one can degrade only the last two

compounds (Junkeret al., 1994).Hydrogenophaga palleronii

S1 has been reported to degrade 4-carbo-4-sulfoazobenzene

by the 4-sulfocatechol pathway via the formation of 4-

aminobenzenesulfate (Vickers, 2002). Another proposed

pathway is transformation of toluene sulfonate to hydroxy

toluene by toluenesulfonate monooxygenase. Pseudomonas

putida strain S-313 catalyzes toluene sulfonate desulfona-tion, which can serve as its sole source of sulfur and leaves 4-

hydroxytoluene unmetabolized. However 4-hydroxy toluene

is a metabolite that is readily catabolized by other bacteria

via the toluene pathway (Eisenmaan & McLeish, 2002).

Another toluene sulfonate degrading bacterium, Coma-

monas testosteroni T-2, was found to contain a degrading

gene on a plasmid (Hooper et al., 1990). Simple alkane

sulfonates are utilized byPseudomonassp. as a carbon source

where crude cell extract catalyzes the oxidation of the




17/44

a-carbon atom of alkanesulfonate to an aldehyde bisulfite

adduct. This adduct then degrades to produce the corre-

sponding aldehyde and sulfite. The substrate range for this

reaction has been reported to be relatively broad where

hydroxy-, methyl-, and alkenyl-substituted compounds are

all transformed (Thysse & Wanders, 1974). Degradation of

alkylsulfate proceeds via initial hydrolysis of the sulfate esterlinkage and subsequent oxidation of the released alkanol

(Kerteszet al., 1994a).Pseudomonassp. C12B and a strain of

Comamonas terrigena were reported to utilize a range of

alkylsulfates as a source of carbon (Payne & Faisal, 1963;

Fitzgerald et al., 1977). Five different alkylsulfatases were

characterized fromPseudomonas sp. C12B and two from C.

terrigena(Dodgson et al., 1982). On the basis of the above

studies, we propose the microbial degradation pathways for

fenamiphos as presented in Fig. 6.

Other organophosphorus pesticides

Several other organophosphorus compounds have been

used extensively for pest control. Diazinon, monocrotophos,

malathion, dimethoate, etc., are being used world-wide.

Several species of bacteria have been isolated and character-

ized that can degrade these compounds in liquid medium

and soils (Table 2).

Fig. 6. Proposed pathways for fenamiphos

degradation by microorganisms. The scheme

is based on articles cited in the text. FSO,

fenamiphos sulfoxide; FSO2, fenamiphos

sulfone; FSO-phenol, fenamiphos sulfoxide

phenol; FSO2-OH, fenamiphos sulfone phenol;

F phenol, fenamiphos phenol.




18/44

Monocrotophos ((3-hydroxy-N-methyl-cis-crotonamide)

dimethyl phosphate) is widely used to control aphids, leaf

hoppers, mites and other foliage pests. It has been classified

as extremely hazardous, with an LD50 value of 20mg kg1

for mammals. The half-life of monocrotophos in soil was

reported to be 4060 days (Tomlin, 2000). Monocrotophos

is easily soluble in water and therefore has potential tocontaminate ground water. Together with its high mamma-

lian toxicity, these characteristics make monocrotophos an

ideal compound for decontamination and detoxification.

Rangaswamy & Venkateswaralu (1992) isolated a monocro-

tophos degrading Bacillus sp. from previously treated soil.

Megharaj et al. (1987) isolated monocrotophos degrading

algae from soil. Two different algae, Aulosira fertilissima

ARM 68 and Nostoc muscorum ARM 221, were found to

utilize monocrotophos as a sole source of phosphorus

(Subramanian et al., 1994). Pseudomonas aeruginosa F10B

and Clavibacter michiganense ssp. insidiosum SBL 11 were

isolated from soil. These bacteria can utilize monocrotophos

as a phosphorus source but not as a carbon source (Singh &

Singh, 2003). Two species ofPseudomonas, three species of

Bacillusand three species of Arthrobacterwere isolated from

soils, which can utilize monocrotophos as a sole source of

carbon (Table 2). Further studies demonstrated that Pseu-

domonas mendocina is the most efficient monocrotophos

degrader among the isolated bacteria and its degrading

capability is plasmid based (Bhadbhade et al., 2002a). The

same group isolated another 17 bacterial isolates from

previously exposed soils which can mineralize monocroto-

phos in liquid culture (Bhadbhade et al., 2002b). The two

most versatile degraders, Bacillus megaterium and A. atro-

cyaneus, were chosen for further studies on the biochemicalmechanisms and pathways of monocrotophos degrada-

tion. Phosphatase activities were observed in both cul-

tures, and it was suggested that the phosphates identified

may be mono- and dimethyl phosphates (Bhadbhadeet al.,

2002b). Dimethyl- and monomethyl phosphates were

involved as intermediates in monocrotophos degradation

in plants and animals (Menzer & Cassida, 1965; Muck,

1994). Another intermediate identified during monocroto-

phos degradation was methylamine, produced by an esterase

enzyme. This esterase could be an amidase capable of

selecting amides as substrates since esterases sometimes

attack the amide bond (Hassal, 1990). Similar pathways ofdegradation were reported for dicrotophos, which is first

demethylated to monocrotophos and then further degraded

to methyl amine (Eto, 1974). As with most of the other

organophosphorus compounds, the first degradation step

of monocrotophos should involve hydrolysis, which

could produce N-methyl acetoacetamide and dimethyl

phosphate (Beynon et al., 1973). Further degradation

of N-methyl acetoacetamide produced valeric acid in A.

atrocyaneusand acetic acid in B. megaterium (Bhadbhade

et al., 2002b). Acetic acid is the key intermediate of the

glycolytic pathway in microorganisms. The pathway of

dicrotophos- and monocrotophos degradation is shown in

Fig. 7.

Degradation of fenitrothion (O,O-dimethylO-4-nitro-m-

tolyl phosphorothioate), a widely used insecticide, byBur-

kholderia sp. strain NF100 was reported (Hayatsu et al.,2000). This strain utilized fenitrothion as a source of carbon

with the help of two plasmids. The first plasmid (pNF2) was

found to catalyze the hydrolysis of fenitrothion to 3-methyl-

4-nitrophenol. The nitro group from this compound

was oxidatively removed to form methylhydroquinone,

which was further metabolized by the second plasmid

(pNF2) (Hayatsu et al., 2000). This bacterium was also

found to degrade p-nitrophenol as a source of energy.

Methylhydroquinone may be degraded by ring fission as

one of the two methods described for p-nitrophenol

degradation in the section dealings with parathion. p-

Nitrophenol degrading Ralstonia sp. SJ98 was reported to

have chemotaxis towards 3-methyl-4-nitrophenol and to

utilize it as a source of carbon. This strain degrades 3-

methyl-4-nitrophenol by the formation of catechol (Bhush-

anet al., 2000b).

Microbial degradation of various other organopho-

sphorus compounds has been documented. Diazinon de-

gradation by a Flavobacterium sp. was reported in 1973

(Sethunathan & Yoshida, 1973). Two Pseudomonas spp.

isolated from sewage sludge were found to degrade diazinon

in a culture medium (Rosenberg & Alexander, 1979). Two

strains of Arthrobacter sp. were reported to hydrolyze

diazinon (Bariket al., 1979). Dimethoate degradation was

reported to be carried out by a plasmid based gene ofP. aeruginosaMCMB-427 (Deshpandeet al., 2001). A novel

dimethoate degrading enzyme was purified and character-

ized from a strain of the fungus Aspergillus niger. This

enzyme was found to degrade all compounds containing

PS linkage like malathion and fermothion but not com-

pounds with the PO linkage (Liuet al., 2001).

Utilization of ethoprophos as a sole source of carbon byP.

putidahas been observed (Karpouzaset al., 2000). Isolation

and metabolism of cadusafos bySphingomonas paucimobilis

and Flavobacterium sp. have been reported recently (Kar-

pouzas et al., 2005). Similarly, several species of bacteria

were isolated from different environments which degradeorganophosphorus compounds in laboratory cultures and

in soils (Singh et al., 1999). Microorganisms isolated from

enrichment of one organophosphorus compound can de-

grade other structurally similar compounds. For example,

Flavobacteriumsp. andP. diminuta were isolated by diazinon

and parathion enrichment but they can degrade a wide

range of other organophosphorus compounds such couma-

phos, me

degradacion de organofosforados

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