DEGRADACION DE PROTEINAS Y AMINOACIDOS. 2 VIAS de DEGRADACION 1. LISOSOMAL 2. PROTEASOMA.
Degradacion de Organofosforados
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Transcript of 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
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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.
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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.
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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 diminuta Co-metabolic Serdaret al. (1982)
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)
Pseudomonas diminuta Co-metabolic Serdaret al. (1982)
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
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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
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435Microbial degradation of organophosphorus compounds
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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.
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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.
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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
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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.
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439Microbial degradation of organophosphorus compounds
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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.
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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
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441Microbial degradation of organophosphorus compounds
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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.
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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
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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.
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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