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Crit Rev Microbiol, Early Online: 1–17! 2014 Informa Healthcare USA, Inc. DOI: 10.3109/1040841X.2014.929564
REVIEW ARTICLE
Microbial degradation of herbicides
Baljinder Singh and Kashmir Singh
Department of Biotechnology, Panjab University, Chandigarh, Punjab, India
Abstract
Herbicides remain the most effective, efficient and economical way to control weeds; and itsmarket continues to grow even with the plethora of generic products. With the development ofherbicide-tolerant crops, use of herbicides is increasing around the world that has resulted insevere contamination of the environment. The strategies are now being developed to cleanthese substances in an economical and eco-friendly manner. In this review, an attempt hasbeen made to pool all the available literature on the biodegradation of key herbicides,clodinafop propargyl, 2,4-dichlorophenoxyacetic acid, atrazine, metolachlor, diuron, glypho-sate, imazapyr, pendimethalin and paraquat under the following objectives: (1) to highlight thegeneral characteristic and mode of action, (2) to enlist toxicity in animals, (3) to poolmicroorganisms capable of degrading herbicides, (4) to discuss the assessment of herbicidesdegradation by efficient microbes, (5) to highlight biodegradation pathways, (6) to discuss themolecular basis of degradation, (7) to enlist the products of herbicides under degradationprocess, (8) to highlight the factors effecting biodegradation of herbicides and (9) to discuss thefuture aspects of herbicides degradation. This review may be useful in developing safer andeconomic microbiological methods for cleanup of soil and water contaminated with suchcompounds.
Keywords
Biodegradation, herbicides, microbes, toxicity
History
Received 3 April 2014Revised 22 May 2014Accepted 27 May 2014Published online 26 August 2014
Introduction
Herbicides are class of chemical compounds that are toxic to
plants, especially unwanted ones. Modern agriculture relies
heavily on herbicides for the control of weeds and ease out to
maximize yield in crops. These compounds have economical
benefits to sustain an increasing world population. The
development of herbicide-resistant plants has also led to an
unexpected increase in the resilience of weeds. Genetically
modified crops resistant to herbicides have become so
prevalent that resistant weeds are beginning to appear,
necessitating new forms of genetic modification. Weeds
have become more and more resistant to herbicides,
prompting farmers to use a wider variety and larger quantity
of herbicides to control them. The introduction of herbicide-
tolerant plants at first decreased herbicide use, but afterwards
increased its usage and scope. While pesticide use dropped
from 22 454 lbs to 15 618 lbs from 2003 to 2008, at a rate of
7000 lbs per acre per year, herbicide use increased from
278 514 000 lbs to 330 422 709 lbs (Cherry, 2010). The world-
wide herbicide market value grew by 39% between 2002 and
2011 and is projected to enhance a further 11% by 2016
(Gianessi, 2013). Majority of herbicides are reported to
constitute between 40 and 60% of pesticides used for
agricultural purpose.
Due to excessive use of herbicides, there is great concern
about their potential environmental hazard. Herbicides con-
tamination can lead to soil and water pollution (Juhler et al.,
2001), reduced biodiversity and depression in soil hetero-
trophic bacteria (including denitrifying bacteria) and fungi
(Song et al., 2013). The environmental fate of herbicides is a
matter of recent concern provided that only a small fraction of
the chemicals reach the target organisms (Pimentel, 1995)
leading to great impacts of residual herbicides in soil and
water on human, animal and crop health. Major sources of
herbicides contamination appear to be an inadequate man-
agement practices specifically involving on-farm handling of
herbicides. The chemical properties and quantity of herbi-
cides determine their toxicity and persistence in the environ-
ment. Their interaction with targeted and nontargeted
organisms has extensively damaged the ecosystem through
entry into the food chains (Singh et al., 2013). According to
the Weed Science Society of America, herbicides have been
classified into 29 different classes based on mechanism of
action (wssa.net/wp-content/uploads/WSSA-Mechanism-of-
Action.pdf). Application method of herbicides generally
include foliar applied method (apply to leaf), soil applied
(with soil contact), broadcast (contact with entire area) and
spot (contact with specified area).
Commonly used herbicides include clodinafop propargyl
(CF), 2,4-dichlorophenoxyacetic acid (2,4-D), atrazine, meto-
lachlor, diuron, glyphosate (GP), imazapyr, pendimethalin,
and paraquat (PQ). These herbicides are harmful to organisms
even at micro (mg) levels. Their uses have resulted in severe
Address for correspondence: Dr. Baljinder Singh, Department ofBiotechnology, Punjab University, Chandigarh, Punjab, India. Tel: +91172 2534076. Fax: +91 172 2541409. E-mail: [email protected],[email protected]
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contamination of the environment, and strategies are now
being developed to clean these substances in an economical
and eco-friendly manner. The present review summarizes
information on the toxicity and microbial degradation path-
ways of these nine herbicides.
Clodinafop propargyl
CF (prop-2-ynyl (R)-2-[(5-chloro-3-fluoro-2-pyridyloxy)
phenoxy]propanoate) is a recently introduced aryloxyphenox-
ypropionate herbicide, used for post emergence control of
annual grasses in cereals (Singh, 2013). The widespread use
of CF has resulted in the discharge of large amounts of the
compound into the environment, which eventually reach the
biosphere. CF is absorbed by the leaves and rapidly
translocated to the growing points of leaves and stems. It
interferes with the production of fatty acids needed for plant
growth in susceptible grassy weeds (Singh, 2013). CF acts by
targeting the enzyme acetyl coenzyme-A-carboxylase, essen-
tial for lipid biosynthesis.
Toxicity of CF
The research focusing on the ecotoxicity of this herbicide is
limited. This might be because of its low persistence and half-
life in soil, which is reported to be 5 d, dependent on the soil
type and pH (Singh, 2013). However, few studies have
demonstrated that CF and its derivatives are toxic and
carcinogenic to humans and other living organisms (Gui
et al., 2011; Kashanian et al., 2008). Kashanian et al. (2008)
reported that the CF in low concentration interacts with calf
thymus DNA by an intercalative mode of binding. CF disrupts
the posterior and ventral development of zebrafish embryos
(Gui et al., 2011; Jaquet et al., 2014). Embryos were exposed
to a range of concentrations from 0.2 mM to 5 mM starting at
late cleavage stage (2 h post fertilization (hpf)) or late
gastrulation stage (10 hpf). The results showed that two
exposure strategies had the same minimum teratogenic con-
centration of 0.6 mM but caused different groups of morpho-
genetic malformations. Thus, due to high toxicity of CF, its
degradation from the environment is of great concern.
Biodegradation of CF
Degradation of CF by aerobic bacteria involves esterase
activity that results in the formation of phenols as metabolites
(Figure 1) (Hou et al., 2011; Singh et al., 2013). Hou et al.
(2011) describe for the first time a microbial strain
Rhodococcus sp. T1, which is able to use CF as a source of
energy. They had reported 97.9% CF degradation without
identifying its metabolites. Singh et al. (2013) isolated
Pseudomonas sp. strain B2 from crop field soil and found
that 87.14% CF was degraded out of initial provided 80 mg/L
CF. The two metabolites clodinafop acid and 4-(4-Chloro-
2-fluoro-phenoxy)-phenol are reported within 9 h of incuba-
tion. Recently, Jaquet et al. (2014) studied the metabolic fate
of 14C-phenyl-labeled CF for 28 d in laboratory assays using
a soil from Germany (Ap horizon, silt loam and cambisol) and
reported that metabolic fate of CF was predominated by the
formation of non-extractable residues (about 60% after 28 d of
incubation). Besides these two studies, no other reports are
available on biodegradation of CF. This might be because of
its low persistence and half life in soil.
2,4-Dichlorophenoxyacetic acid
2,4-D is one of the most widely used chlorinated acidic
phenoxy herbicide in the world and is an analogue of a growth
hormone, auxin. 2,4-D was first synthesized in 1941,
marketed commercially and registered for use in the United
States in 1944 and 1948, respectively. It has provided
economical, selective, effective control of broadleaf weeds
in agriculture crops, pastures and forests for the past decades.
This is used both agriculturally and domestically for post-
emergent control of broad-leaf weeds. 2,4-D is formulated as
amine salts (mainly dimethyl-amine salt), which are more
soluble in water than acid, and ester derivatives (2-ethyhexyl
ester), which are readily dissolved in an organic solvent. The
2,4-D amine salts and esters rapidly convert to the 2,4-D acid
under most environmental conditions (US Environmental
Protection Agency (EPA), 2004).
Figure 1. CF degradation by microbes (1) CF(2) acid metabolite, clodinafop acid; (3) 4-(4-Chloro-2-fluoro-phenoxy)-phenol (4) phenol.The scheme is based on articles cited in thetext.
N
Cl
F
O
O
CH3
CH3
CO2 H2O
O
O
CH
N
Cl
F
O
O
O
OH
N
Cl
F
O
OH
OHN
Cl
F
OH
+ +
1
2
3
4
2 B. Singh Crit Rev Microbiol, Early Online: 1–17
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In plants, 2,4-D functions by maintaining high levels of the
plant hormone auxin, resulting in overstimulation of plant
growth and ultimately death. 2,4-D also induced ethylene
production and therefore, it acts as defoliating agent. After
application, 2,4-D is readily absorbed by the leaves and roots,
and is then translocated to all parts of the plant. However, the
exact mechanism by which this herbicide affects cells is not
completely understood. It was reported that 2,4-D is a
peroxisome proliferators and especially in plant cells, the
herbicide induced mitotic and meiotic abnormalities both in
vivo and in vitro (Gonzalez et al., 2005; Marron et al., 2006).
2,4-D is a moderately persistent chemical with a half- life
between 20 and 200 d. Weintraub et al. (1954) studied the
metabolism and persistence of 2,4-D in dormant plant tissue
and reported relatively long persistence of the compound
in bean plants by performing experiments with C14-labeled
2,4-D. The persistence of 2,4-D in soil depends on the type of
soil. It is highly water-soluble acids (pKa¼ 3.11) and have a
low tendency to accumulate in organic matter (El-Bestawy &
Hans-Jorgen, 2007), therefore can enter as contaminants into
streams, rivers or lakes directly from drainage of agricultural
lands (Lagana et al., 2002; Mikov et al., 2010; Muller &
Babel, 2004). In the aqueous environment, 2,4-D is most
commonly found as the free anion (Burns et al., 2012).
Toxicity of 2,4-D
2,4-D is classified by WHO as a hormonal herbicide of level
II toxicity. Exposure to human beings may occur through
inhalation, skin contact or ingestion. In most cases, the
predominant route of occupational exposure has been the
absorption of spills or aerosol droplets through the skin
(Chaudhry & Huang, 1988). 2,4-D is easily adsorbed into the
human organ from the alimentary tract and skin and is
subsequently excreted in the urine in nearly unchanged form.
It was detected in stomach, blood, brain and kidney of four-
day-old rat neonates fed by 2,4-D-exposed mothers (Cenkci
et al., 2010; Sturtz et al., 2000). 2,4-D has been shown to
exert toxic effects on both animals and humans such as
carcinogenic, teratogenic, neurotoxic, immunosuppressive,
cytotoxic and hepatoxic effects. The toxicity of 2,4-D have
been studied and reviewed by many workers (Atamaniuk,
2013; Bortolozzi et al., 2001; Bukowska, 2006; Burns et al.,
2012; Coady et al., 2013; Garabrant & Philbert, 2002; Mikov
et al., 2010; Munro et al., 1992). In mammals, 2,4-D disrupts
energy production (adenosine triphosphate (ATP)) (Palmeira
et al., 1994). 2,4-D is associated with Hodgkin’s disease,
non-Hodgkin’s lymphoma (NHL) and soft tissue sarcoma
(Garabrant & Philbert, 2002; Holland et al., 2002).
In mammalian cells in vitro, 2,4-D inhibits cell growth and
protein synthesis by affecting the DNA (Gonzalez et al.,
2005). Ganguli et al. (2014) have investigated the molecular
mechanism of 2,4-D-induced lung toxicity in A549 and WI38
cell lines and reported IC50 values 126 ± 2.25 mM and
115 ± 4.39 mM, in A549 and WI38 cells, respectively, for 72 h.
In plant, it causes an increase in chromosomal aberration,
multipolar cells and a decrease in mitotic index (Gui et al.,
2011). The genotoxic potential of 2,4-D in plant has been
investigated by using different test systems including micro-
nucleus, comet assay and Random Amplified Polymorphic
DNA (Aksaka et al., 2013; Cenkci et al., 2010).
2,4-D is among the most used herbicides in cultivable
lands and its high toxicity in both animals and plants reinforce
researched to found out the degradation pathways along with
genes involved in such pathways.
Biodegradation of 2,4-D
Several bacterial strains have been described that are able to
use 2,4-D as the sole carbon and energy source (Baelum et al.,
2010; Marron et al., 2006; Sandoval-Carrasco et al., 2013).
The most commonly cited 2, 4-D degrading genera are
Pseudomonas, Alcaligenes, Ralstonia, Delftia, Arthrobacter
and Burkholderia. Degradation of 2,4-D via oxidative cleavage
of ether bond with subsequent chlorophenol hydroxylation
followed by the modified ortho-cleavage pathway of chlor-
ocatechols has been demonstrated for most of these isolates
(Figure 2). The electron effects, the spatial orientation and the
hydrophobic effect of their substituent groups have obvious
influence on their degradation pathway. Evans et al. (1971)
isolated two strains of Pseudomonas sp. capable of degrading
2,4-D and proposed biodegradation pathway of 2,4-D. Plasmid
Cl
Cl
OO
O-
Cl
OO
O-
Cl
HO
Cl
HO
OH
TCA
Cl
Cl
HO
Cl
HO
Cl
OHClO
HO
ClO
O TCA
4Chlorophenoxyacetate
4-chlorophenol
4-chlorocatechol
3,5-DCC 2,4-DC cis muconate2,4-DCP
2,4-D
intradiol
ring cleavage
intradiol
ring cleavage
2,4-dehalogenase
monooxygenase hydroxyalase
tfdA
ketoglutarate deoxygenase
tfdB
hydroxyalasetfdC
1,2-dioxygenase
tfdDEFKR
oxoglutarate
succinate
NADPH
NADH+
Figure 2. Degradation of 2,4 D by microbes. 2,4-DCP, (2,4-dichlorophenol), 3,5-DCC (3.5-Dichlorocatechol) TCA, tricarboxylic acid cycle. Thescheme is based on articles cited in the text.
DOI: 10.3109/1040841X.2014.929564 Microbial degradation of herbicides 3
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involvement in the degradation of 2,4-D was first reported by
Fisher et al. (1978). They reported that ability to degrade 2,4-
D is encoded by a 58-megadalton conjugal plasmid, pJPl.
Genes encoding 2,4-D degradation are often located on
conjugative plasmids, pEMT1 and pJP4 (Bhat et al., 1994;
Chong & Chang, 2009; Don & Pemberton, 1981; Top et al.,
1995), but have been found to be also chromosomally located
in Burkholderia spp (Matheson et al., 1996). Plasmid pEMT1
contains the same degradative genes in an organization similar
to that of plasmid pJP4 but does not belong to the same
incompatibility group as pJP4. Don & Pemberton (1985)
constructed a genetic and biophysical map of pJP4 by using
transposon mutagenesis, deletion analysis, gene cloning and
restriction analyses. Plasmid pJP4 contains the genes for the
degradation of 2, 4-D to chloromaleyl acetic acid, whereas
chromosomal genes of the host are necessary for complete
mineralization of the compound. The modified ortho cleavage
pathway in degradation of 2,4-D is encoded on conjugative
plasmids (Bhat et al., 1994; Top et al., 1995; Poh et al., 2002;
Garabrant & Philbert, 2002), which give information on
movement of genes by horizontal gene transfer. The genes
responsible for 2,4-D degradation pathway (tfdA, -B, -C,-D, E
and F) are localized to the transmissible plasmid, pJP4 is well
understood and the enzymes participating in the pathway have
been purified and characterized (Amy et al., 1985; Annelie
et al., 2001; Atamaniuk et al., 2013; Evans et al., 1971). The
tfd genes encoding these enzymes have been localized, cloned
and sequenced (Don et al., 1985; Hoffmann et al., 2003;
Tiedje et al., 1969; Tsutsui et al., 2013). The regulatory
mechanisms of tfd gene expression have been elucidated
(Inoue et al., 2012). Kitagawa et al. (2002) cloned and
characterized new family of 2,4-D degradation genes,
cadRABKC, from Bradyrhizobium sp. strain HW13. The
cadR gene was inferred to encode an AraC/XylS type of
transcriptional regulator from its deduced amino acid
sequence. The cadABC genes were predicted to encode 2,4-
D oxygenase subunits from their deduced amino acid
sequences. The cadK gene was presumed to encode a 2,4-D
transport protein from its deduced amino acid sequence that
showed 60% identity with the 2,4-D transporter, TfdK, of strain
Ralstonia eutropha JMP134. Gene bioaugmentation with
conjugative plasmids-harboring bacteria capable of degrading
2,4-D on the indigenous microbial community has been
evaluated by several researchers previously (Dejonghe et al.,
2000; Don & Pemberton, 1981; Inoue et al., 2012; Newby
et al., 2000; Tsutsui et al., 2013). Hoffmann et al. (2003)
reported that the putative genes of the complete 2,4-D
degradation pathway are organized in a single genomic unit
in alkalitolerant strain Delftia acidovorans P4a.
Atrazine
Atrazine, [2-chloro-4-(ethylamino)-6-(isopropylamino)-s-tria-
zine], is a selective herbicide belonging to the family of the s-
triazines. Because of its high mobility in soil and its massive
application, atrazine has often been detected in surface and
ground waters at concentrations well above the permissible
limits (Biradar & Rayburn, 1995; Hayes et al., 2002; Kolpin &
Kalkhoff, 1993; Tappe et al., 2002). The moderate persistent,
chemical water solubility (33 mg l�1 at 20 �C) and low soil
sorption partition coefficient (Kd¼ 3.7 L kg�1) are key factors
influencing its potential contamination to aquifers, ground-
water and rain water as it normally percolates to groundwater
or river via infiltration (Siripattanakul et al., 2009). Atrazine
inhibits photosynthesis and its associated noncyclic photopho-
sphorylation is in higher plants (Shelton et al., 1996). All
higher plants probably metabolize atrazine by N-dealkylation
with some species such as corn and wheat, which contain
benzoxazinone, and utilize hydroxylation as well.
Toxicity of atrazine
The main target of atrazine in humans and animals is the
endocrine system. This herbicide is associated with relatively
high chronic toxicity. One study showed the presence of14C-labeled atrazine in soil 22 years after the last application
(Jablonowski et al., 2008), revealing the risk of chronic
atrazine exposure. Atrazine is a carcinogen that affects the
central nervous system, reproductive system, immune system
and cardiovascular function. Exposure to atrazine in animals
is associated with some types of NHL in adult humans. In
2007, the USEPA began reviewing several epidemiological
cancer studies concerning atrazine and its possible association
with carcinogenic effect in humans. Atrazine is a potential
disruptor of normal sexual development in male frogs and
also alter some aspects of the immune response (Yaw-Jian
et al., 1999). Studies in rodents suggested that atrazine causes
alterations in locomotor activity (Bardullas et al., 2011),
decreases striatal catecholamine content and decreases the
number of tyrosine hydroxylase-positive neurons (TH+) in the
substantia nigra pars compacta and ventral tegmental area
(Hayes et al., 2002). Due to ubiquitous and unpreventable
water contamination, the European Union announced a ban of
atrazine in October 2003, but same month, the USEPA
approved continued use of atrazine.
Due to high toxicity at ppb level, bioremediation in
atrazine contaminated soil and water is necessary and
received many attentions around the world.
Biodegradation of atrazine
Microbial metabolism has been regarded as the most important
mechanism of atrazine degradation in soil. A number of
microorganisms with different atrazine degradation efficien-
cies and growth characteristics have been reported (Table 1).
One of the well-known atrazine-degrading bacteria is
Pseudomonas sp. strain ADP, which can decompose atrazine
to NH3 and CO2 by some enzymes encoded by six genes,
atzABCDEF (Siripattanakul et al., 2009). Another best-
characterized strain whose atrazine metabolic track is different
from Pseudomonas sp. strain ADP is Arthrobacter aurescens
TC-1. It can metabolize atrazine to cyanuric acid through a
degrading pathway catalyzed by TrzN, AtzB and AtzC
enzymes (Figure 3). In addition, there are also some
researchers reported the hybrid pathways involving the trzDN
or atzABCDEF genes combinations found in atrazine-degrad-
ing strains or consortiums (Hess & Warren, 2002). To date, the
known genes that encode the enzymes to hydrolyze atrazine are
atzA, atzB, atzC, atzD, atzE, atzF, trzD and trzN. Although
these genes have their own substrate specificity and are
harbored in different atrazine-degrading bacteria isolated from
geographically distinct locations, nearly all of the genes are
4 B. Singh Crit Rev Microbiol, Early Online: 1–17
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Tab
le1
.R
epo
rted
mic
roo
rgan
ism
sca
pab
leo
fd
egra
din
gco
mm
on
lyu
sed
her
bic
ides
.
Atr
azin
e
Mic
roo
rgan
ism
Iso
lati
on
fro
mo
rso
urc
eD
egra
dat
ion
pat
hw
ayG
ene
invo
lved
ind
egra
dat
ion
pat
hw
ayR
efer
ence
sA
gro
ba
cter
ium
rad
iob
act
erJ1
4a,
So
ilM
etab
oli
tes
pro
du
ced
are
hy
dro
xyat
ra-
zin
e,d
eeth
yla
traz
ine
and
dee
thylh
yd
rox
yat
razi
ne
–S
tru
ther
set
al.,
19
98
Kle
bsi
ella
sp.
A1
and
Co
ma
mo
na
ssp
.S
ewag
eo
fa
pes
tici
de
mil
lH
yd
rox
yat
razi
ne,
N-i
sop
rop
yla
mm
elid
ean
dcy
anu
ric
acid
atz
A,
B,
C,
D,
E,
Fgen
ew
ere
clo
ned
Yan
get
al.,
20
10
Art
hro
ba
cter
sp.
stra
inD
NS
10
,B
lack
soil
–a
tzA
,a
tzD
,a
tzE
,a
tzF
and
trzD
trzN
,a
tzB
and
atz
Cgen
ew
ere
clo
ned
Zh
ang
etal
.,2
01
1
Rh
od
oco
ccu
ssp
.st
rain
MB
-P1
So
ilD
e-et
hyla
traz
ine
and
de-
iso
pro
pyla
traz
ine
Pla
smid
-en
cod
edF
azlu
rrah
man
etal
.,2
00
9A
rth
rob
act
ersp
.st
rain
HB
-5In
du
stri
alw
aste
wat
erH
yd
rox
yat
razi
ne
and
cyan
uri
cac
id–
Wan
get
al.,
20
11
Pse
ud
om
on
as
stra
inY
AY
A6
Gar
den
soil
Deg
rad
atio
np
roce
eded
via
dec
hlo
rin
atio
nan
dN
-dea
lkyla
tio
n–
Yan
ze-K
on
tch
ou
&G
schw
ind
,1
99
4A
rth
rob
act
ersp
.st
rain
DA
T1
So
ilD
egra
dat
ion
gen
estr
zN,
atz
Ban
da
tzC
on
pla
smid
DN
AW
ang
etal
.,2
01
3
Art
hro
ba
cter
sp.
TE
S6,
Agri
cult
ura
lso
iltr
zN,
atzB
,an
dat
zC.
Seb
aıet
al.,
20
11
Clo
din
afo
p-p
rop
arg
yl
(CF
)P
seu
do
mo
na
ssp
.st
rain
B2
Cro
pfi
eld
soil
.E
ster
ase
acti
vit
y:
stra
inB
2d
egra
ded
CF
tocl
od
inaf
op
acid
and
4-(
4-c
hlo
ro-2
-fl
uo
ro-p
hen
ox
y)-
ph
eno
lw
ith
in9
h.
–S
ing
het
al.,
20
13
Rh
od
oco
ccu
sw
rati
sla
vien
sis
So
ilC
lod
inaf
op
acid
–Ja
qu
etet
al.,
20
14
2,4
-DB
rad
yrh
izo
biu
msp
.st
rain
HW
13
Bu
ried
soil
-2
,4-D
deg
rad
atio
ngen
es,
cad
RA
BK
CK
itag
awa
etal
.,2
00
2S
erra
tia
ma
rces
cen
san
dP
enic
illi
um
spB
razi
lian
soil
con
tam
inat
edw
ith
2,4
-Dh
erb
icid
e2
,4-d
ich
loro
ph
eno
l(2
,4-D
CF
)–
Sil
va
etal
.,2
00
7
Alc
ali
gen
eseu
tro
ph
us
So
il2
,4-d
ich
loro
ph
eno
l2
,4-D
mo
noxy
gen
ase
Per
kin
s&
Lu
rqu
inet
al.,
19
88
Ra
lsto
nia
eutr
op
ha
JMP
13
4H
awai
ian
soil
Ph
eno
ld
eriv
ativ
esca
dR
AB
KC
Kit
agaw
aet
al.,
20
02
Met
ola
chlo
rF
usa
riu
msp
.,M
.ra
cem
osu
sM
eto
lach
lor-
trea
ted
farm
soil
,2
-ch
loro
-N-(
2-e
thy
l-6
-hy
dro
xy
met
hylp
he-
nyl)
-N
-(2
-met
hox
y-1
-met
hyle
thy
l)-
acet
amid
e
–S
axen
aet
al.,
19
87
Azo
tob
act
ersp
.st
rain
SS
B8
1.
–4
-ch
loro
ph
eno
lan
d4
-ch
loro
cate
cho
l,O
xid
ativ
ep
ath
way
Gau
riet
al.,
20
12
Gly
ph
osa
teO
chro
ba
ctru
ma
nth
rop
iG
PK
3,
Ach
rom
ob
act
ersp
.S
od
-po
dzo
lso
ilH
igh
deg
rad
atio
nca
pac
ity
wit
ho
ut
the
accu
mu
lati
on
of
any
met
abo
lite
,in
par
-ti
cula
r,am
ino
met
hylp
ho
sph
on
icac
id
–E
rmak
ova
etal
.,2
01
0
Tri
cho
der
ma
viri
de
stra
inF
RP
3F
ore
stso
il–
–A
rfar
ita
etal
.,2
01
3F
usa
riu
mst
rain
s9
11
48
(Fu
sari
um
oxy
spo
rum
)Is
ola
ted
fro
msu
gar
can
eA
min
om
eth
ilp
ho
sph
on
icac
id(A
MP
A).
–C
astr
oJr
etal
.,2
00
7
Pse
ud
om
on
assp
.st
rain
PG
29
82
.G
lyp
ho
sate
pro
cess
was
test
ream
Pri
mar
ym
etab
oli
team
ino
met
hylp
ho
-sp
ho
nat
ew
aso
bse
rved
,5
%o
fth
eg
ly-
ph
osa
tew
asd
egra
ded
by
ase
par
ate
pat
hw
ayin
vo
lvin
gb
reak
dow
no
fg
lyp
ho
-sa
teto
gly
cin
e.
–Ja
cob
etal
.,1
99
8
Fla
vo
bac
teri
um
spec
ies.
Ind
ust
rial
acti
vat
edsl
ud
ge
AM
PA
–D
iuro
nA
rth
rob
act
erst
rain
sD
47
Bri
tish
and
Fre
nch
agri
cult
ura
lso
ils
Am
mo
nia
and
CO
2.
–T
urn
bu
llet
al.,
20
01
3,4
-dic
hlo
roan
ilin
e,C
O2
.–
Sø
ren
sen
etal
.,2
00
8
(co
nti
nu
ed)
DOI: 10.3109/1040841X.2014.929564 Microbial degradation of herbicides 5
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nly.
generally highly conserved and plasmid borne. Furthermore,
associating with transposon, some of the genes could widely
spread between bacteria regardless of their genera. This may
enable more microbes or consortiums to mobilize atrazine and
mobilizes some new catabolic pathways.
The degradation of atrazine occurs predominantly by
biological processes, including N-dealkylation, dechlorination
and ring cleavage. Atrazine biodegradation can be initiated by
N-dealkylation of the ethyl or isopropyl side chains to
produce deethylatrazine (DEA) or deisopropylatrazine (DIA).
Dechlorination has been reported as an early step in atrazine
metabolism, and two different s-triazine hydrolase enzymes
have been characterized. In some microorganisms, complete
biodegradation of atrazine to ammonia and CO2 has been
obtained. Pseudomonas sp. strain ADP might be the best-
characterized atrazine mineralizing one (Mandelbaum et al.,
1993). The genes encoding the three enzymes that are
responsible for the conversion of atrazine to cyanuric acid
were atzA, B and C. The research on atrazine-degrading
microorganisms has been directed to the isolation and
characterization of natural occurrence lineages in environ-
ments contaminated with this pesticide. For the purpose of
potential bioremediation practice, a large variety of atrazine-
degrading bacteria from diverse genera have been isolated
(Rousseaux et al., 2001). Among bacteria, there are reports
on atrazine degradation by individual strains such as
Pseudomonas sp. (Mandelbaum et al., 1995), Rhodococcus
rhodochrous, Acinetobacter spp., Agrobacterium sp.,
Microbacterium sp., Bacillus sp., Micrococcus sp.,
Deinococcus sp. and D. acidovorans (Vargha et al., 2005),
as well as by species consortia including Agrobacterium
tumefaciens, Caulobacter and Pseudomonas sp. ADP is now
the most known and the best-characterized atrazine-degrading
bacterium (Wackett et al., 2002). However, microorganisms
from genus Arthrobacter are well known for their strong
capacity to degrade atrazine and have been isolated from
agricultural and heavily contaminated soils at spill sites and
industrial wastewaters from atrazine production plants (Zhou
et al., 2012). The atrazine-degrading bacteria generally initi-
ate the degradation through a hydrolytic dechlorination,
catalyzed by the enzyme atrazine chlorohydrolase (AtzA),
encoded by the atzA gene, followed by two hydrolytic
deamination reactions catalyzed by hydroxyatrazine ethylami-
nohydrolase (AtzB) and N-isopropylammelide isopropyl-
amino-hydrolase (AtzC), encoded by the genes atzB (trzB) e
atzC (trzC), respectively (De Souza et al., 1998), which
convert atrazine sequentially to cyanuric acid that is then
completely mineralized to CO2 and NH3 by other three
hydrolases. In some bacterial strains, the biodegradation of
atrazine initiate through N dealkylation of the lateral ethyl and
isopropyl chains to DEA and DIA (Kaufman & Blake, 1970).
Pseudomonas sp. ADP is the best-characterized bacterial strain
capable to degrading the herbicide atrazine. The atrazine
catabolic pathway in this bacterium contains six enzymatic
steps encoded by atzABC and the atzDEF genes. The atzABC
genes have been shown to be widespread and plasmid borne in
a number of bacteria isolates (De Souza et al., 1998:
Rousseaux et al., 2001; Topp et al., 2000; Wackett et al.,
2002). In Pseudomonas sp. ADP, the atzABCDE genes are
harbored on the catabolic plasmid pADP-1 (Martinez et al.,
Tab
le1
.C
on
tin
ued
Atr
azin
e
A.
glo
bif
orm
isst
rain
D4
7an
dV
ari
ovo
rax
sp.
SR
S1
6C
on
tam
inat
edso
ilan
dw
ater
reso
urc
esB
.b
ass
ian
a,
C.
eleg
an
s,A
gri
cult
ura
lan
dn
on
-ag
ricu
ltu
ral
soil
s3
,4-d
ich
loro
anil
ine,
Am
mo
nia
and
CO
2.
pu
hA
gen
eC
asti
llo
etal
.,2
00
6
Mic
roco
ccu
ssp
.st
rain
PS
-1D
iuro
nst
ora
ge
site
Deg
rad
atio
np
roce
eded
via
de-
alk
yla
tio
nan
dfi
nal
lysy
nth
esiz
ing
amin
od
eriv
a-ti
zed
met
abo
lite
–S
har
ma
&S
uri
,2
01
1
Par
aquat
Asp
erg
illu
sn
iger
span
dP
seu
do
mo
na
ssp
UK
san
dy
loam
soil
sC
arb
on
dio
xid
eas
the
ult
imat
em
iner
al-
izat
ion
pro
du
ct–
Ric
ket
tes
etal
.,1
99
9
Lip
om
yces
sta
rkey
iC
lay
soil
Ace
tic
acid
and
carb
on
dio
xid
e–
Car
ret
al.,
19
86
Pen
dim
ath
alin
Fu
sari
um
oxy
spo
rum
,A
sper
gil
lus
ory
zae,
Len
tin
ula
edo
des
,P
enic
illi
um
bre
vico
mp
act
um
and
Lec
an
icil
liu
msa
ksen
ae
So
ilD
eset
hy
l-te
rbu
thyla
zin
e–
Pin
toet
al.,
20
12
6 B. Singh Crit Rev Microbiol, Early Online: 1–17
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iew
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h on
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rson
al u
se o
nly.
2001). In Pseudomonas sp. ADP, the atzDEF operon encodes
cyanuric acid amidohydrolase (AtzD), biuret amidohydrolase
(AtzE) and allophanate hydrolase (AtzF), involved in cleavage
of the cyanuric acid to carbon dioxide and ammonia, which is
assimilated as a nitrogen source (De Souza et al., 1998).
Garcia-Gonzalez et al. (2003) have demonstrated that nitrogen
control of atrazine metabolism is functional under soil
conditions and may therefore limit the potential of
Pseudomonas sp. strain ADP for atrazine bioremediation in
nitrogen-fertilized agricultural soils. The atzABC genes are
constitutively expressed and are not regulated either by
induction of atrazine or by repression of other N sources in
this strain (Martinez et al., 2001; Devers et al., 2007). The
atzDEF genes are divergently transcribed from AtzR, predicted
N
N
N
Cl
HN NHCH3
NH2
NH2
NH2
H3C
H2N
H2N
H2N
H3C
H2N
CH3
H3C CH3
CH3
CH3
CH3
CH3
H3C CH3
N
N
N
OH
NH NH
Atrazine
Hydroxyatrazine
N
N
N
OH
NH OH
N-isopropylammelide
N
N
N
OH
O HOH
Cyanuric acid
N
N
N
Cl
HN
deethylatrazine
N
N
N
Cl
NH
deisopropylatrazine
N
N
N
Cl
deisopropyl deethylatrazine
N
N
N
Cl
OH
N
N
N
OH
OH
ammelide
N
N
N
Cl
O HNH
N-ethylammelide
NH
OH
O O
+NHO O
BiuretAllophanate
AC
HAEA
IAIA
AM DEAM
TH
HAEA
IAIA
TC
EAA
CAH
BH AH
2-chloro 4-hydroxy 6-amino 1,3,5-triazine
atzA
atzB
atzC
atzD
atz E atzF
NH3
NH2NH2
CO2
Figure 3. Atrazine biodegradation along with gene and enzymes involved in degradation pathways: AC, atrazine chlorohydrolase; HAEA,hidroxyatrazine ethylaminohydrolase; IAIA, N-isopropylammelide isopropylamidohydrolase; TC, s-triazine chlorohydrolase; AM, atrazinemonooxygenase; DEAM, deethylatrazine monooxygenase; DIHA, deisopropyhidroxylatrazine amidohydrolase; EAA, N-ethylammelide amidohy-drolase, TH, s-triazine hydrolase; CAH, cyanuric acid hydrolase; BH, biuret hydrolase; and AH, allophanate hydrolase. The scheme is based onarticles cited in the text.
DOI: 10.3109/1040841X.2014.929564 Microbial degradation of herbicides 7
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to encode a transcriptional LysR type regulator (LTTR). A
putative LTTR-binding site can in fact be found upstream of
atzD gene, thereby suggesting that transcription of the atzDEF
operon may be regulated and the protein encoded by the orf99
(AtzR) play a role in this regulation. The atzDEF operon
resides in a contiguous cluster adjacent to the orf99, a potential
transcriptional LTTR (Martinez et al., 2001). Atrazine bio-
degradation can be positively or natively affected by the
external addition of organic carbon sources (Abdelhafid et al.,
2000) and is dependent on types of added carbon sources found
that the addition of citrate to soils resulted in enhanced atrazine
degradation by Arthrobacter sp. strain KU001. A good
candidate for bioaugmentation is the Arthrobacter strain
DAT1 that has shown high atrazine-degrading efficiency
both in liquid cultures and soils (Wang & Xie, 2012). This
strain can utilize atrazine as a sole nitrogen source for growth
and harbors the atrazine-metabolic genes trzN, atzB and atzC.
The cited literatures on biodegradation of atrazine help us
to know the metabolic ability and gene characteristics of
strains involved in biodegradation of atrazine. However, there
is need to study strains evolution by comparing the metabolic
ability and characterizing genes involved in atrazine
biodegradation.
Metolachlor
Metolachlor (2-chloro-N-(2-ethyl-6-methylphenyl)-N-(meth-
oxyprop-2-yl)acetamide) is a selective chloroacetamide herbi-
cide used to control broadleaf and annual grass weeds in corn
(Zea mays L.), soybean (Glycine max L. Merr.), peanut
(Arachis hypogaea L.) and potato (Solanum tuberosum L.)
(Gaynor et al., 1993; Thurman et al., 1992). Large quantities
of metolachlor (22 million kg of active ingredient) are applied
to agricultural fields in the United States, particularly in the
Midwest, where most of the corn and soybeans are grown.
Metolachlor is a member of the chloroacetanilide herbicide
chemical family. Other members include acetochlor, alachlor,
butachlor, butenachlor, delachlor, diethatyl, dimethachlor,
metazachlor, propachlor, propisochlor, prynachlor, terbuchlor,
thenylchlor. When it is absorbed through the roots and shoots, it
acts as a growth inhibitor by suppressing synthesis of chloro-
phyll, proteins, fatty acids and lipids, isoprenoids (including
gibberellins) and flavonoids (including anthocyanins).
Toxicity of metolachlor
Studies have shown that metolachlor affects cell growth, and its
low levels cause cytotoxic effects in lymphocytes along with a
significant decrease in mitotic index (Rollof et al., 1992).
Studies with HepG2 cells, an immortalized human liver cell
line, exposed to 100 ppb metolachlor for 24 h had 21% fewer
numbers of cells compared to nontreated control cells (Hartnett
et al., 2013). Recently, Lowry et al. (2013) studied decrease
of HepG2 cell growth after metolachlor exposure and reported
that the levels of the retinoblastoma protein including two of its
hyperphosphorylated forms are decreased in metolachlor
exposed cells possibly leading to cell cycle arrest.
Biodegradation of metolachlor
A major breakdown pathway of metolachlor in the soil
is by both aerobic and anaerobic microorganisms. The
transformation by soil microorganisms of metolachlor to its
primary degradates: metolachlor ethane sulfonic acid (ESA)
and metolachlor oxanilic acid (OA) (Figure 4) has been
suggested to occur as a result of shifting of chlorine atom of the
parent compound by glutathione, followed by the formation of
the ESA and OA degradates by different enzymatic pathways
(Barbash et al., 1999). Degradation of metolachlor in soil
occurs mainly by microbial decomposition (Xu et al., 2008)
and photo-degradation. Microbial degradation rates are
affected by soil depth, organic carbon and dissolved oxygen
concentrations, temperature and size of microbial populations.
In case of sandy soils, a half-life of 81 d for anaerobic microbial
populations and 67 d for aerobic microbial populations were
reported in the laboratory. Photo-degradation occurs only when
metolachlor is present on the soil surface. Fifty percent of
surface-applied metolachlor can be degraded in eight days on
soil, while only 6% degrades over one month in soils where
metolachlor was incorporated into the surface layer.
Diuron
Diuron (3,4-dichlorophenyl)-1,1-dimethylurea) is a systemic
substituted wide-spectrum phenylurea herbicide used for
weed control in agricultural crops and non crops areas
(Castillo, 2006; Stasinakis et al., 2009). Non crops areas
include along fence lines, pipelines, powerlines, railway lines,
roads, footpaths; in timber yards and storage areas; and
around commercial, industrial and farm buildings, electrical
substations and petroleum storage tanks. Diuron is used as an
algaecide in ornamental ponds, fountains and aquaria and
mildewcide in paints. Diuron is often used either alone or in
combination of other herbicides such as bromacil, hexazi-
none, PQ, thiadiazuron, imazapyr, monosodium, sodium
chlorate, sodium metaborate and copper sulfate (USEPA,
2004). Diuron is available in wettable powder, granular,
flowable, pelleted/tableted, liquid suspension and soluble
concentrate formulations. Technical grade (the grade that is
usually used for agricultural purposes) diuron is a white,
crystalline and odorless solid.
Diuron is a systemic substituted phenylurea herbicide. It is
easily taken up from soil solution through the root system of
plants and rapidly translocated into stems and leaves by the
transpiration system, moving mainly via the xylem. Diuron
primarily functions by blocking the Hill reaction in
Metolachlor [2-chloro-N-ethylphenyl)-N-(2-methoxy-1-methylethyl) acetamide]
(2-ethyl-6-methylphenyl)(2-methoxy-1-methylethyl)amino oxo-acetic acid
Metolachlorethane sulfonic acid
Metolachlor oxanilic acid
morpholinone + Carbinol
CO2 + H2O
Figure 4. Biodegadation of metolachlor. The scheme is based on articlescited in the text.
8 B. Singh Crit Rev Microbiol, Early Online: 1–17
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photosynthesis process, limiting the production of high-
energy compounds such as ATP, which is used for several
metabolic processes. Diuron binds to the QB-binding site on
D1 protein of the photosystem-II complex in chloroplast
(thylakoid membranes), thus blocking electron transport from
QA to QB. This process prevents CO2 fixation and the
production of ATP and other high energy compounds, which
are needed for plant growth. The inability to reoxidize QA
promotes the formation of triplet state chlorophyll, which
interacts with ground state oxygen to form singlet oxygen.
Both triplet chlorophyll and singlet oxygen can extract
hydrogen from unsaturated lipids, producing a lipid radical
and initiating a chain reaction of lipid per-oxidation. Lipids
and proteins are attacked and oxidized, resulting in loss of
chlorophyll and carotenoids, and in leaky membranes, which
cause cells and cell organelles to dry and disintegrate rapidly
(Hess & Warren, 2002).
Toxicity of diuron
Diuron is highly toxic to aquatic organisms, LC50 (48 h)
values for diuron range from 4.3 to 42 mg/L in fish, and from
1 to 2.5 mg/L for aquatic invertebrates/organisms. The LC50
(96 h) is 3.5 mg/L for rainbow trout (EXTOXNET et al.,
1996). Residues of diuron was detected in milk, fat body,
muscles, liver and kidney of cows, cows fed with very low
doses of diuron in their diets (EXTOXNET et al., 1996).
Domingues et al. (2011) studied diuron exposure in male
Wister rats and they found that diuron exerts systemic and
target-organ toxicity, mainly at higher concentration. This
research records the potential mutagenic, teratogenic, repro-
ductive and carcinogenic effects of diuron (Federico et al.,
2011) investigated the genotoxic effects of chlorotoluron,
diuron and difenoxuron by analyzing chromosomal and sister
chromatid exchange in exposed mammalian cells (Chinese
hamster). They found that phenylurea herbicides induced
direct genotoxic activity, but the cytogenetic effects were
greatly enhanced after metabolic conversion.
Biodegradation of diuron
Diuron is susceptible to degradation by soil microorganisms,
and enriched cultures of aquatic microorganisms from pond
water could also degrade diuron to 3,4-dichloroaniline as a
major metabolite (Figure 5) (Fratila-Apachitei et al., 1999).
Bogaerts et al. (2000) studied microbial degradation of
diuron and ecotoxicology to investigate its breakdown after
application to soils. Quantitative biodegradation assays were
executed with fungal strains, showing that diuron was
degraded but not entirely. A series of tests were taken out
to select the most efficient fungal strain for diuron degrad-
ation. Among the fungal strains, only four strains were able to
transform diuron to an extent (up to 50%) after seven days of
incubation: Beauveria bassiana, Caenorhabditis elegans,
Phanerochaete chrysosporium and Mordellistena isabellina.
Diuron degradation by the fungal strains led to the formation
of two metabolites obtained in different proportions according
to the microorganism. For the fungal strains, diuron degrad-
ation led to the formation of the demethylated products. The
identified metabolites were synthesized in sufficient amounts
to confirm their structures and determine their non-target
toxicity using four biotests. According to the Microtox test,
the metabolites N-(3,4-dichlorophenyl)-N-methylurea and
N-3, 4-dichlorophenylurea presented a three times higher
toxicity than that of diuron. Dalton et al. (1966) reported that
the removal of N-methyl groups eliminates herbicidal activity
of diuron. Decomposition is followed by the removal of the
urea group, which results in the formation of 3,4-dichloroani-
line, ammonia and CO2.
Glyphosate
GP [N-phosphonomethyl glycine] is an active ingredient of
herbicides applied to annual and perennial weeds. GP inhibits
5-enolpyruvilshikimate-3-phosphate synthase, the enzyme of
shikimic acid synthesis that participates in the synthesis of
aromatic amino acids. GP is a broad-spectrum, non-selective
systemic herbicide that can be used to control most weeds,
both annual and perennial plants, under many varied
situations such as agriculture, forestry, orchards, vineyards,
industry and no-till cropping systems and has been domes-
ticity classified as an easily degradable herbicide in the past
(Celis et al., 2008). Although GP is believed to be a relatively
safe compound, in our days, reports can be found that GP has
negative effects on human health. In addition, nowadays, GP
and one of its principal metabolites (aminomethylphosphonic
acid) have frequently been detected in ground water (Laitinen
et al., 2007; Meza-Joya et al., 2013). The application of GP
results in the yellowing and decay of leaves within 5–10 d
Figure 5. Biodegradation of diuron: DCPU[1-(3,4-dichlorophenyl) urea], DCPMU[3-(3,4-dichlorophenyl)-1-methylurea].
Cl
Cl NH
N
O
Diuron
Cl
Cl NH
NH
O
DCPMU
Cl
Cl NH O
DCPU
Cl
Cl
dichloroaniline
Cl
NH O
monuron
NH2
NH2
NH2
H3C CH3CH3
DOI: 10.3109/1040841X.2014.929564 Microbial degradation of herbicides 9
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(sometimes 30 d) caused by the breakdown of aromatic amino
acids synthesis. Penaloza-Vazquez et al. (1995) reported that
GP remains unchanged in the soil for varying lengths of time,
because of its adsorption on clay particles and organic matter
present in the soil. This condition makes this herbicide very
persistent in soils and sediments.
Toxicity of GP
Application of GP in the agriculture for a long period of time
increases the susceptibility of crops to diseases and in arable
areas causes dieback in hedgerow trees. In soil, GP molecules
may be adsorbed on humus and may also form complexes
with metal cations (Fe2+ Cu2+, Mn2+ and Ni2+) (Veiga et al.,
2001). Soil amendment with P fertilizers – sources of
phosphate ions, which compete with GP for binding sites –
leads to accumulation of free GP, available for plant root
uptake and for microbial metabolism (Simonsen et al., 2008).
GP absorbed into the plant not only mainly through its foliage
but also through soft stalk tissue. It is then translocated to
growing points of the plant where it acts on various enzyme
systems inhibiting aromatic amino acids: tyrosine, tryptophan
and phenylalanine that are essential for protein formation and
secondary products in susceptible plants. This pathway also
works in higher plants and microorganisms but not in animals.
However, GP-containing products are acutely toxic to animals
(Cox, 1995). Plants treated with GP slowly die over a period
of days or weeks, and because the chemical is transported
throughout the plant, no part survives. GP is chosen for early-
season weed control before planting and after harvesting
(Duke & Powles, 2008; Laitinen, 2007). In reduced tillage or
no-till cultivations, GP is used to prepare fields before
planting, during crop development and post harvest.
Biodegradation of GP
GP is an amphoteric and non-volatile compound, no
photodegradation happens and it is stable in air. It is
practically insoluble in most of organic solvents, for instance,
ethanol, acetone and benzene, because of its high polarity, but
it is completely soluble in water (Laitinen et al., 2007). The
degradation of GP is slower in soils with a higher adsorption
capacity. Degradation rate was also affected by the particular
microbial community of each soil (Carlisle & Trevors, 1986).
Microorganisms known to degrade GP include Pseudomonas
sp., Arthrobacter atrocyaneus and Flavobacterium sp.
(Table 1). The primarily metabolite of GP is aminomethyl-
phosphonic acid (Figure 6), which is non-toxic and degraded
by microbes at a somewhat slower rate than the parent
compound (Rueppel et al., 1977; Carlisle & Trevors, 1986).
Imazapyr
Imazapyr [2-(4,5-dihydro-4-methyl-4-(1-methylethyl)-5-oxo-
1H-imidazol-2-yl)-3-pyridinecarboxylic acid] also known as
Chopper, Arsenal and Assault is a non-selective broad-
spectrum systemic herbicide used for control of annual and
perennial grasses, broadleaf branch, sedge weeds and a
variety of shrubs and deciduous trees. It was first registered in
the United States in 1984. Imazapyr is absorbed by the foliage
and roots, with rapid transfer to the xylem and phloem to the
meristematic regions, where it accumulates and causes
disruption of protein synthesis and interferes with cell
growth and DNA synthesis. The result of exposure is death
of new leaves. Imazapyr comes in two forms: weak acid and
isopropylamine salt (49% water solution), although commer-
cial products are almost solely the isopropylamine salt
form. Imazapyr is persistent in the soil with the half-life of
17 months (USEPA, 1984).
Toxicity of Imazapyr
Imazapyr is considered as non-toxic to animals; however, it is
irritating to the eyes and can cause rashes, redness and
swelling at the site of exposure. It stops the biosynthesis of
essential chemical (aliphatic three branched-chain amino
acids) in plants. Animals also require these chemicals but they
do not synthesize it and take directly from plants. Therefore,
specific toxic affect can only happen in a plant, and toxic
affects in animals occur at a much higher dose.
Biodegradation of imazapyr
The herbicide is commonly found in two forms: weak acid and
isopropylamine salt (Figure 7), although commercial products
are almost solely the isopropylamine salt. Biodegradation of
imazapyr by Streptomyces sp. strain PSI/5 was first observed
by Shelton et al. (1996). Xuedong et al. (2005) isolated two
Figure 6. Biodegradation of glyphosate. Thescheme is based on articles cited in the text.
Arthrobacter atrocyaneusEnterobacter aerogenes
P
HO
O
NH
O
OH
NH
O
OH
GlyphosateSarcosine
P
HO OH
OH
OOMAD
FormaldehydeMethlyamineAMPA
C-P lyase
C-P lyase
Intermediary Metabolism
C1 metabolic cycle
GD
Geobacillus caldoxylosilyticusFlavobacterium sp.
NH2NH2
H2CH3C
H3C
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bacterial species, Pseudomonas and Bacillus, from soil that
were capable of degrading imazapyr. They found degradation
rate of more than 70% at 50 or 100 ppm concentration of
initially added imazapyr to the culture after 48 h of incubation.
Further studies are needed to study the mechanism of
biodegradation of imazapyr by microbes.
Pendimethalin
Pendimethalin (N-(1-ethylpropyl)-3, 4-dimethyl-2, 6-dinitro-
benzenenamine, a dinitroaniline herbicide) is both a pre-
emergence and early post-emergence herbicide, used for
control of most annual grasses and many annual broad-leaved
weeds in crop fields. Pendimethalin was first registered as a
pesticide in the United States in 1972. Pendimethalin has
relatively long persistence in soil due to immobilization with
low leaching potential or due to hydrophobic nature of
pendimethalin adsorb strongly to organic matter and clay
minerals. It degrades more rapidly in anaerobic soil than in
aerobic soil conditions (Megadi et al., 2010). Its mode of
action is inhibition of mitotic cell division in developing root
systems.
Toxicity of pendimethalin
The USEPA has classified pendimethalin as persistent
bioaccumulatice toxic; it is of low acute toxicity, but causes
thyroid follicular cell adenoma. It is highly toxic to fish and
aquatic invertebrates. It is moderately persistent in aerobic
soil environments. It is therefore important to investigate the
degradation mechanism of pendimethalin.
Biodegradation of pendimethalin
Biodegradation of pendimethalin in soil was reported by
several investigators under both aerobic and anaerobic
conditions with bacteria and fungi in different types of soils
(Zheng & Cooper, 1996). Four metabolites were formed and
identified as N-(1-ethylpropyl)-3,4-dicarboxy 2,6-dinitroben-
zenamine-N-oxide, N-(1-ethylpropyl)-3,4-dimethoxy-2,6-
dinitrobenzenamine and benezimadazole-7-carboxyaldehyde
(Figure 8). The reactions involved were monohydrolysis of 2-
methyl groups followed by dihydrolysis. Further oxidation of
amine groups and hydroxylation of propyl groups produced
the above-mentioned metabolites.
Paraquat
PQ (1,1-dimethyl-4,4-bipyridinium) is a quaternary nitrogen
herbicide widely used for control of broadleaf weed. It is
nonselective compound that destroys green plant tissue on
contact by disrupting photosynthesis and rupturing cell
membranes, which allows water to escape leading to rapid
desiccation of foliage. PQ has strong affinity to bound to clay
minerals and organic matter in the soil.
Toxicity of PQ
PQ is toxic to living organisms including human beings as it is
easily absorbed practically by all routes including the gastro-
intestinal tract, skin, mucous membranes and lungs (Huang
et al., 2012). PQ poisoning is a major medical problem in Asia
where it is used for suicides in rural communities (Wua et al.,
2013). The acute toxic effects of PQ in the lung, liver and
kidney have been well recognized in animals (Huang et al.,
2012; Wua et al., 2013). In toxicity studies using laboratory
animals, PQ has been shown to be highly toxic by the inhalation
route and has been placed in Toxicity Category I (the highest of
four levels) for acute inhalation effects.
Baldwin et al. (1966) studied toxicity of PQ and reported
that PQ caused deterioration in the lungs. A dermal toxicity
study using rabbits resulted in scabbing and inflammation
when tested at the two highest doses (2.6 mg cation/kg group
and 6.0 mg cation/kg group). In an inhalation toxicity study,
rats were exposed to respirable aerosols (particle size less
than 2 um in diameter) of PQ dichloride, which resulted
in lung changes and extensive sores and swelling in the
larynx.
Biodegradation of PQ
PQ is considered a toxic to soil fungi and bacteria causing a
reduction in their population (Sahid et al., 1992). Several
bacterial and fungal isolates obtained from soil and
Figure 7. Biodegradation of imazapyr.
N
O
OH
N
NH
O
N
O
O−
N
NH
O
NH+
Imazapyr acid Imazapyr isopropylamine salt
Pyridine hydroxy-dicarboxylic acid Nicotinic acid
Pyridine dicarboxylic acid (quinolinic acid)
degraded into three possible product
CH3CH3
CH3
CH3CH3
CH3
CH3H3C
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wastewater can degrade PQ. The contact herbicide PQ is used
to control weeds in a wide range of crops. When PQ enters the
soil environment, it is rapidly and strongly bound to clay
minerals and organic matter and deactivated. Various studies
have used pure cultures of soil micro-organisms to elucidate
the degradative pathways of ring-labeled PQ. These studies
established the range of bacteria and fungi able to degrade
PQ (e.g. Corynebacterium fascians Dows, Lipomyces starkeyi
Loo and Rij, Aspergillus niger van Teigh, Penicillium
frequentans West, Fusarium sp and Pseudomonas sp) and
that conditions in soil solution are conducive to the degrad-
ation of PQ. In majority of the cases, however, PQ degrad-
ation was shown to be extremely variable and evidence of
only one degradation product other than CO2 has been
reported (Figure 9), the N-methyl betaine of isonicotinic acid
(Funderburk & Bozarth, 1967). The latter product has,
however, been shown to be a major intermediate in the
photolytic degradation of PQ (Slade, 1965).
NHO
O
O
O
NHO
+
Pendimethalin
6 -Aminopendimethalin
3,4-dimethyl -2,6-dinitroaniline Pentane
dealkylation
nitroreduction
CH3
NHO
O
OH
NHO
O
OOH
NHO
O
NH
N
1 2
3
4
O−
O−
O−
O−O−
O−
O−
O−
O−
O−
O−
N+N+
N+N+
N+
N+ N+
N+ N+
N+ N+
H3C
CH3
CH3
CH3
CH3
CH3
CH3 CH3
CH3
CH3CH3
CH3
CH3
CH3
CH3
CH3
CH3
H2N
CH3
CH3
CH3CH3
Figure 8. Biodegradation of pendimathalin. Four metabolites were formed and identified as N-(1-ethylpropyl)-3,4-dicarboxy 2,6-dinitrobenzenamine-N-oxide, N-(1-ethylpropyl)-3,4-dimethoxy-2,6-dinitrobenzenamine and benezimadazole-7-carboxyaldehyde. The reactions involved weremonohydrolysis of 2-methyl groups followed by dihydrolysis. Further oxidation of amine groups and hydroxylation of propyl groups produced theabove-mentioned metabolites.
12 B. Singh Crit Rev Microbiol, Early Online: 1–17
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Factors effecting microbial degradation of herbicides
Based on cited literature, it has been found that the inoculum
size, amounts of additional co-substrates carbon and nitrogen
compounds, organic matter of soil and pH are the major
factors that affected the extent and rate of herbicides
degradation.
Agricultural soils are rich in nitrogen due to routine
fertilization, and most atrazine-degrading bacteria use atra-
zine as a nitrogen source. However, Yang et al. (2010)
reported members of Klebsiella sp. A1 and Comamonas sp.
A2 capable of degrading atrazine and are insensitive to
exogenous nitrogen sources. Wang et al. (2013) reported
addition of both carbon and nitrogen sources promotes
degradation rate of atrazine. Struthers et al. (1998) reported
that addition of Agrobacterium radiobacter J14a cells in soil
resulted in two to five times higher mineralization of atrazine
than in the noninoculated soil. However, sucrose addition did
not result in significantly faster mineralization rates or shorten
degradation lag times.
The mineralization of GP in soils is individually regulated
and correlated by exchangeable H+, soil pH, oxalate extract-
able Al3+ and bacterial cell numbers. Recently, Al-Rajab &
Hakami (2014) studied GP degradation in three agriculture
soil and reported rapid degradation with a half-life of 14.5 d in
the silt clay loam soil incubated at 20 �C.
Johnson & Sims (2011) studied solvent toxicity in soil for
bioavailability of 2,4-D toward microorganisms and suggested
that solvent toxicity should be balanced with uniformity of
substrate distribution when using organic carriers in soils.
Conclusion and future directions in biodegradationof herbicides
Due to excessive use of herbicides, ecosystems are under
threat of its pollution. Microbes are the main vehicle for
remediation of herbicides, and new discoveries, such as novel
biodegradation pathways, multispecies interactions and com-
munity-level responses to herbicides addition, are helping us
to understand, predict and monitor the fate of herbicides. With
the recent release of new metagenomic information from
herbicides-associated environments determining the micro-
bial players offers the promise of new discovery. Genomic,
transcriptomic and mutant studies conducted with defined co-
cultures are yielding new information about how microbes
interact and benefit energetically via different mechanisms of
interspecies electron transfer. Despite this, there are many
challenges, not least because of the heterogeneity of these
ecosystems and the structure of herbicides. For example, there
is growing awareness about the toxicity of the herbicides,
which are difficult to degrade. Microbial metabolism is a
process of energy conversion, and it is governed by enzymatic
mechanisms, where reaction intermediates play a vital role.
Screening of organisms that degrade herbicides or produce
enzymes or enzyme systems that degrade herbicides may
prove as environmentally profitable in the present time.
Figure 9. Biodegradation of paraquat. M(microorganisms) – L. starkeyi,Corynebacterium fascians, Lipomyces star-keyi, Aspergillus niger, Penicillium frequen-tans, Fusarium sp and Pseudomonas sp.1¼ 1,10-dimethyl-4,40-bipyridinium, 2¼ 1-methyl-4,40-bipyridinium, 3¼ 1,10-dimethyl-2-oxo2,3-dihydro-4,4-bipyridinium, 4¼N-methyl-isonicotinic acid or 4-carboxyl-1-methyl-pyridylium, 5¼ succinic acid, 6¼oxalic acid, 7¼ formic acid and8¼methylamine.
O
N
O
OH
O
OH
O
OHO
OHO
OH O
OH
H +++
++
+
1
23
4
5 6 7 8
hv, Mhv
hv, Mhv
M
+hv
H3C H3C
H3CH3C
H3C
NH2 CO2
CH3
CH3
N+
N+N+N+
N+N+
H3C NH2 CO2
H2O CO2NH3
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A screening program for such organisms and enzymes is
required but will require more universally uniform standards
for assessment of their degradative ability. Current research
provides an understanding of how the evolution of promis-
cuous enzymes and the recruitment of enzymes available from
the metagenome allows for the assembly of biodegradation
pathways. Nevertheless, physicochemical constraints includ-
ing bioavailability, bioaccessibility and the structural vari-
ations of similar chemicals limit the evolution of
biodegradation pathways.
Further study is required in genetic modification of
microorganisms. The use of gene probes for studying the
distribution of set of genes in herbicides contaminated soils
will be useful in identifying niches in which these kinds of
genes prevail and the conditions under which the population
of microbes bearing these genes increases (Singh et al.,
2013). Phytoremediation in conjunction with rhizospheric
microbes may provide a cheap, fast, eco-friendly and efficient
rhizoremediation processes for the removal of explosive waste
from the upper layers of the soil (Singh et al., 2012). The
implementation of advanced technologies such as proteomics
and bioinformatics should be investigated to provide more
knowledge regarding the enzymatic mechanisms and inter-
mediates involved in metabolic activities during biodegrad-
ation. In this regard, research may be also focused on
examining the possibility of using the enzymes rather than the
microorganisms in biological treatment. To sum it up, the
synergistic performance of various microorganisms and
various technologies is to be considered a research topic of
high priority.
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