FEMS Microbiology Ecology 45 (2003) 1^11
www.fems-microbiology.org
MiniReview
Microbial degradation of isoproturon and related phenylurea
herbicides in and below agricultural ¢elds
Sebastian R. Srensen
a
a;
, Gary D. Bending b , Carsten S. Jacobsen a , Allan Walker b ,
Jens Aamand a
Department of Geochemistry, Geological Survey of Denmark and Greenland, ster Voldgade 10, DK-1350 Copenhagen, Denmark
b
Horticulture Research International, Wellesbourne, Warwick, UK
Received 12 February 2003 ; received in revised form 11 April 2003 ; accepted 18 April 2003
First published online 17 May 2003
Abstract
The phenylurea herbicides are an important group of pesticides used extensively for pre- or post-emergence weed control in cotton, fruit
and cereal crops worldwide. The detection of phenylurea herbicides and their metabolites in surface and ground waters has raised the
awareness of the important role played by agricultural soils in determining water quality. The degradation of phenylurea herbicides
following application to agricultural fields is predominantly microbial. However, evidence suggests a slow degradation of the phenyl ring,
and substantial spatial heterogeneity in the distribution of active degradative populations, which is a key factor determining patterns of
leaching losses from agricultural fields. This review summarises current knowledge on the microbial metabolism of isoproturon and
related phenylurea herbicides in and below agricultural soils. It addresses topics such as microbial degradation of phenylurea herbicides in
soil and subsurface environments, characteristics of known phenylurea-degrading soil micro-organisms, and similarities between metabolic
pathways for different phenylurea herbicides. Finally, recent studies in which molecular and microbiological techniques have been used to
provide insight into the in situ microbial metabolism of isoproturon within an agricultural field will be discussed.
3 2003 Federation of European Microbiological Societies. Published by Elsevier Science B.V. All rights reserved.
Keywords : Biodegradation; Degradation pathway ; Spatial heterogeneity; Linuron; Diuron; Sphingomonas spp.
1. Introduction
Since the discovery and marketing of phenylurea herbicides shortly after the Second World War, this group has
grown to be one of the most important classes of herbicides used worldwide. The phenylurea herbicides are
mostly used either pre- or post-emergence in cotton, fruit
or cereal production. However, diuron (Table 1) is also
used as a total herbicide in urban areas and as an algicide
in antifouling paints. The majority of the currently available phenylurea herbicides are either N-methoxy-N-methyl- (e.g. linuron and chlorobromuron) or N,N-dimethylsubstituted (e.g. isoproturon and diuron) compounds. A
selection of the most commonly used phenylurea herbicides is presented in Table 1.
The phenylurea herbicides generally have relatively high
water solubilities and low tendencies to sorb to soil, ren-
* Corresponding author. Tel. : +45 3814 2317 ; Fax: +45 3814 2050.
E-mail address : srs@geus.dk (S.R. Srensen).
dering them mobile in soil. Several phenylurea herbicides
and their metabolites have been detected as contaminants
of groundwater [1,2], rivers and streams [1,3], lakes [4,5]
and seawater [6] in di¡erent parts of the world. Phenylurea
herbicides may enter the environment in several di¡erent
ways, for example by point source contamination as a
consequence of accidental spillage or from washing of
spraying equipment. However, most of the herbicide
load enters the environment as di¡use contamination following normal spraying practice (e.g. [7,8]). After introduction into agricultural soils, several processes a¡ect the
vertical and horizontal distribution of the herbicides including transport by water £ow, sorption to soil components and various degradation processes. Degradation can
involve biotic and abiotic processes, where microbially facilitated biodegradation is especially interesting, as it is a
major process in the complete mineralisation of aromatic
compounds to harmless inorganic products [9]. The natural attenuation rate of phenylurea herbicides, with respect
to mineralisation of the phenyl structure to CO2 , is either
not detectable or very slow in samples from groundwater
0168-6496 / 03 / $22.00 3 2003 Federation of European Microbiological Societies. Published by Elsevier Science B.V. All rights reserved.
doi:10.1016/S0168-6496(03)00127-2
FEMSEC 1524 2-6-03
2
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
Table 1
Molecular structure of common phenylurea herbicides. Compare to
Fig. 1 for identi¢cation of known metabolites for each of the herbicides.
waters in 2000, with 14% of the analysed samples exceeding 0.1 Wg l31 [13], the limit concentration for individual
pesticides in surface and ground waters serving as drinking
water set by the European Community Drinking Water
Directive. IPU was furthermore among the seven most
frequently encountered pesticides in a survey of both marine waters and groundwater in the UK in 2000 [13].
In this review we summarise the current knowledge on
the microbial metabolism of IPU and related phenylurea
herbicides in and below agricultural soils. We consider
topics such as the characteristics of known phenylureadegrading soil micro-organisms and similarities between
metabolic pathways for di¡erent phenylurea herbicides.
Finally, we focus on recent studies which have applied
molecular and microbiological techniques to provide insights into the in situ microbial metabolism of IPU within
a British agricultural ¢eld.
2. Major degradation processes for phenylurea herbicides in
agricultural soils
The phenylurea herbicides presented in Table 1 are stable to chemical degradation in aqueous solution under
I.
B
II.
B
D
O
A
N
H
D
O
N
N
CH3
A
N
H
CH3
1.
B
O
NH
A
aquifers [10^12]. The degradation of phenylurea herbicides
in plough layer soils is thus of major interest because most
agricultural ¢elds function as recharge zones for aquifers,
rivers and lakes, and thereby serve as biological ¢lters
determining the degree to which the herbicides biodegrade
before transport by water.
The herbicide isoproturon (IPU), used against annual
grasses and broad-leaved weeds, is among the most extensively used pesticides in European cereal production. Approximately 3300 tonnes was applied on 3.0 million hectares of agricultural land in the UK in 1997, making IPU
the most widely used organic pesticide in the UK [13]. Due
to widespread and extensive usage, IPU has been detected
as a contaminant of rivers, streams, lakes, marine waters
and groundwater across Europe [2,3,5,6,14]. In the UK,
IPU was the most frequently detected pesticide in surface
N
H
CH3
4.
2.
B
O
NH2
A
N
H
3.
B
A
B
NH2
A
NH2
Fig. 1. Proposed general degradation pathways for N-methoxy-N-methyl- and N,N-dimethyl-substituted phenylurea herbicides in agricultural
soil. Pathway I: Involving sequential N-dealkylations (steps 1 and 2)
and hydrolysis to aniline derivatives (step 3). Pathway II: Direct hydrolysis to the aniline derivatives (step 4). See Table 1 for identi¢cation of
substituents A, B and D for each of the phenylurea herbicides and their
metabolites.
FEMSEC 1524 2-6-03
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
moderate temperatures and within a pH range of 4^10
(e.g. [5,15,16]), which suggests that chemical degradation
is of minor importance in most agricultural soils. Studies
have however shown that photochemical degradation may
occur when the herbicides are exposed to sunlight (e.g.
[5,17]). These photochemical processes lead to only partial
degradation, creating products that may accumulate in the
environment. Such an accumulation is not desirable, as
evidence suggests that some of these degradation products
(Fig. 1, steps 1^3) may be more hazardous to non-target
organisms than is the herbicide itself [18^20]. Some of
these products have additionally been shown to persist
and contribute to contamination of surface and ground
water [1,4,10,21]. A study examining the ecotoxicity of
IPU and three of its metabolites created by both photochemical and microbial processes (Fig. 2, steps 1^3) revealed a 50% reduction in methanogenesis in pond sediment following exposure to metabolite concentrations
ranging from 400 to 550 Wg l31 [18]. In contrast, below
10% reduction was observed with the highest IPU concentration of 6250 Wg l31 . Similar increases in the toxicity
following initial degradation of diuron, chlorotoluron
and IPU have been reported with toxicity assays towards
bacteria and ciliated protozoan populations [19,20].
Although photochemical processes could be involved in
the degradation of phenylurea herbicides on soil or plant
surfaces, evidence suggests that the degradation in agricultural soils occurs predominantly through the activities of
soil micro-organisms, as shown for IPU [22], diuron [23],
£uometuron [24], linuron, chlorobromuron and metobromuron [25,26]. Besides being the major mechanism for
degradation of phenylurea herbicides in soils, microbiological processes also constitute the ultimate steps in the mineralisation of the phenyl ring, as herbicide metabolites
may enter the general microbial metabolism leading to
production of CO2 and biomass (e.g. [27^29]).
3. Microbial attenuation of phenylurea herbicides in soil
and subsurface environments
3.1. Degradation of phenylurea herbicides in the soil
environment ^ ¢eld studies
A ¢eld experiment in the UK by Gooddy et al. [8] revealed high concentrations of diuron and some of its
known metabolites under the plough layer of a calcareous
soil following normal ¢eld treatment in a monitoring programme lasting for 50 days. Diuron was detected in porewater from depths of 9^54 cm in concentrations greater
than 1 Wg l31 in 97% of the measurements made. Three
metabolites known from initial degradation of diuron
(Fig. 1, steps 1^3) were also detected in soil and porewater
with concentrations exceeding 1 Wg l31 throughout the
entire monitoring period [8]. Another recent ¢eld study
in the UK demonstrated that after normal agricultural
3
practice, IPU and chlorotoluron may persist in soil and
that leaching to the underlying aquifer may occur for at
least 3 years after ¢eld treatment [7]. Occurrence of metabolites was not monitored in this study. Schuelein et al.
[21] detected IPU and a range of di¡erent metabolites in
soil, porewater, drainage pipes and in a nearby creek following treatment of a German agricultural ¢eld with IPU.
This study reported a rapid leaching and surface runo¡ of
IPU and its metabolites concomitant with a heavy precipitation event 14 days after ¢eld application.
Taken together the ¢eld studies indicate a slow natural
degradation rate of the three herbicides diuron, IPU and
chlorotoluron following agricultural ¢eld treatments resulting in leaching losses after subsequent rain events.
However, ¢eld experiments alone are inadequate for detailed studies of microbial degradation processes and for
clarifying the mechanisms underlying the potential for microbial attenuation of phenylurea herbicides in soil and
subsurface environments.
3.2. Microbial degradation and bioavailability of
phenylurea herbicides in agricultural soils
Several laboratory studies have indicated a slow natural
biodegradation rate of phenylurea herbicides in various
soils and subsurface environments (e.g. [10,24,28,30^32]).
Typically, in laboratory microcosm experiments with agricultural soils, 5^25% of added 14 C-phenyl-labelled IPU
(14 C-IPU) is mineralised to 14 CO2 within 2^3 months at
about 20‡C [11,30,33^35]. However, recent studies have
shown a rapid and extensive mineralisation of IPU in
some previously ¢eld-treated soils, with 40^50% 14 C-IPU
being metabolised to 14 CO2 within 1 month at 15^20‡C
[29,36,37], suggesting an in situ microbial adaptation to
IPU metabolism following repeated application at the
same ¢eld. There is a limited amount of information on
the degradation of the phenyl ring of other phenylurea
herbicides, but generally it seems that the halogenated
compounds, e.g. diuron, linuron and £uometuron, are
very slowly mineralised in agricultural soils [24,31,32]. Zablotowicz et al. [32] reported a mineralisation of 14 C-phenyl-labelled £uometuron to 14 CO2 of less than 3% in agricultural soils over 25 days of incubation at 28‡C. Berger
[31] compared the mineralisation of di¡erent 14 C-phenyllabelled phenylurea herbicides, including linuron, metobromuron, chlorotoluron and IPU, in three arable soils
but found no production of 14 CO2 within 56 days at
20‡C or 30‡C.
Large fractions of 14 C-labelled residues not extractable
by various organic solvents have been observed following
treatment of soil with 14 C-IPU or other 14 C-labelled phenylurea herbicides in numerous laboratory studies. Between 25 and 75% of the initially applied 14 C-IPU can
be converted to non-extractable residues following 2^3
months of incubation [30,33^36,38]. It has been suggested
that the metabolite 4-isopropyl-aniline (4IA) produced by
FEMSEC 1524 2-6-03
4
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
O
4.
6.
N
N
H
H3 C
O
CH3
H3 C
HO
H3 C
N
H
1-OH-IPU
7.
O
O
H
N
H3 C
N
H
H3 C
H3 C
MDIPU
N
H
2-OH-IPU
11.
O
O
H3 C
H3 C
NH2
HO
N
H
CH3
2-OH-MDIPU
DDIPU
O
H
N
N
H
H3 C
[N-(4-(1-hydroxy-1-methylethyl)phenyl)
-N'-methylurea]
[N-(4-isopropylphenyl)urea]
O
CH3
[N-(4-(1-hydroxy-1-methylethyl)phenyl)
-N',N'-dimethylurea]
8.
2.
CH3
N
HO
CH3
[N-(4-isopropylphenyl)-N'-methylurea]
H3 C
CH3
[N-(4-(2-hydroxy-1-methylethyl)phenyl)
-N',N'-dimethylurea]
1.
H3 C
CH3
N
CH3
IPU
5.
H2 C
9.
3.
12.
O
H3 C
HN
H3 C
H3 C
H3 C
14.
CH3
H3 C
4IA
[4-isopropyl-aniline]
NH2
HO
NH2
NH
10.
N
H
2-OH-DDIPU
[N-(4-(1-hydroxy-1-methylethyl)phenyl)urea]
13.
H3 C
15.
CH3
[4,5-bis-(4-isopropylphenylamino)
-3,5-cyclohexadiene-1,2-dione]
H3 C
HO
H3 C
CH3
H3 C
N
?
N
NH2
2-OH-4IA
[4-(1-hydroxy-1-methylethyl)aniline]
CH3
H3 C
Mineralisation
[4,4-diisopropyl-azobenzene]
Fig. 2. Proposed degradation pathways for IPU in agricultural soil and by de¢ned soil micro-organisms. Compounds shown in boxes are dead-end metabolites without any further degradation. See text for details (Sections 5.1 and 5.2).
partial degradation of IPU (Fig. 2, steps 3^5) may be
irreversibly bound to components of the soil organic matter [35]. This binding signi¢cantly decreases the mineralisation of the phenyl structure in soil systems [35,39], due
to the lower availability of 4IA to degrading soil microorganisms [40]. Detection of large fractions of non-extractable 14 C-labelled residues following soil treatment with
other 14 C-ring-labelled phenylurea herbicides has also
been reported (e.g. [32]). The similarities between the metabolic pathways for phenylurea herbicides, as outlined in
Fig. 1, suggest that these observations could also be related to the binding of aniline metabolites to soil components. The possible mechanisms involved in the binding of
aniline compounds to soil components have been reviewed
by Parris [41]. In contrast to the aniline metabolites, the
phenylurea herbicides themselves sorb to a lesser extent
(e.g. [40]), and it seems unlikely that the initial degradation
of these compounds should be limited by sorption in agricultural soils shortly after ¢eld application. However, prolonged contact times between the herbicide and soil components may increase the fraction unavailable for microorganisms [40,42], and this might reduce the microbial
metabolism of phenylurea herbicides with ageing of resi-
dues in the soil environment. This ageing phenomenon in
soil is not fully understood. Suggested mechanisms include
di¡usion of pesticide into organic matter matrices or micropores where it is less accessible to potential degrading
micro-organisms [40,42].
3.3. Biodegradation of phenylurea herbicides in the
subsurface environment
Despite the frequent detection of phenylurea herbicides
in groundwater, studies on the fate of these compounds in
subsurface environments are scarce. The potential for
biodegradation of IPU decreases below the plough layer
[10^12], and no reports of signi¢cant mineralisation of
the phenyl structure of IPU, or of any other phenylurea
herbicides, in samples from subsurface environments have
so far been presented. Johnson et al. [43] observed a potential for partial biodegradation of IPU resulting in accumulation of the metabolite N-(4-isopropylphenyl)-NPmethylurea (MDIPU) (Fig. 2, step 1) in samples from a
chalk aquifer, but no further degradation was detected.
Occurrence of polar metabolites was not monitored in
this study, and the activity of alternative metabolic path-
FEMSEC 1524 2-6-03
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
ways involving hydroxylation (Fig. 2, steps 6^8) can therefore not be excluded.
Many groundwater aquifers have anoxic zones with
metabolic activity based on electron acceptors other than
oxygen. The potential biodegradation of phenylurea herbicides under anoxic conditions is largely unknown. Larsen et al. [11] found no mineralisation of IPU anaerobically
in the presence of nitrate in microcosm experiments with a
sandy aquifer sediment. A later study failed to detect any
mineralisation of IPU in di¡erent aquifer sediments under
denitrifying, sulfate-reducing or methanogenic conditions
following incubation for 312 days at 10‡C [44]. Detection
of metabolites from partial degradation of IPU was not
included in these studies. Previous studies with di¡erent
phenylurea herbicides including diuron and fenuron (Table 1) revealed a potential for reductive dehalogenation of
diuron in anoxic pond sediments [45]. No degradation of
fenuron was observed under anoxic conditions. However,
it remains unknown if a potential for partial degradation
of phenylurea herbicides exists in anoxic subsurface environments.
The lack of extensive phenylurea degradation in samples
obtained below the plough layer could be related to several parameters, such as unfavourable growth conditions
for indigenous microbial degraders, or alternatively the
lack of competent degradative microbial populations.
However, factors determining the presence of active phenylurea-metabolising populations in subsurface environments still remain to be clari¢ed. The long residence times
of pesticides in groundwater may suggest that even very
slow degradation rates could be of importance in determining future drinking water quality.
4. Phenylurea-degrading soil micro-organisms
4.1. Characterisation of de¢ned microbial consortia able to
degrade phenylurea herbicides
The detection of IPU and other phenylurea herbicides
as water contaminants and the apparent low microbial
degradation potential in soil and subsurface environments
have stimulated studies aimed at characterising natural
microbial populations able to metabolise these compounds. Enrichment-culture techniques have been used
with varied success in attempts to enrich and isolate phenylurea-degrading micro-organisms. In several studies,
slurries of mineral media and agricultural soils failed to
extensively degrade IPU [28,31,34]. Lehr et al. [34] used
two enrichment strategies involving inoculation of four
di¡erent agricultural soils into a de¢ned mineral medium
with IPU as sole carbon source, or with IPU and glucose
as an alternative carbon source. No extensive degradation
of IPU was observed in any of the cultures, indicating a
lack of enrichment during this procedure. Trace amounts
of the metabolites MDIPU, N-(4-isopropylphenyl)urea
5
(DDIPU), N-(4-(2-hydroxy-1-methylethyl)phenyl)-NP,NPdimethylurea (1-OH-IPU) and N-(4-(1-hydroxy-1-methylethyl)phenyl)-NP,NP-dimethylurea (2-OH-IPU) (Fig. 2)
were detected in the culture media [34]. Berger [31] attempted to enrich micro-organisms from two agricultural
soils using 10 di¡erent phenylurea herbicides individually,
including IPU, diuron, monuron, chlorotoluron, fenuron,
metobromuron, chlorobromuron and linuron as sole
source of carbon and nitrogen, but these attempts were
also unsuccessful. Srensen and Aamand [28] enriched a
mixed bacterial culture growing on the IPU metabolites
MDIPU and 4IA as sole carbon sources, whereas no enrichment cultures were obtained from the same agricultural soil with IPU as the sole carbon source. The lack of
success in obtaining enrichment cultures with phenylurea
herbicides in these studies could be related to the soils
used, which in many cases were agricultural soils without
previous exposure to the herbicides. The phenylurea-degrading enrichment cultures presented in the literature so
far are all derived from soils previously treated with phenylurea herbicides.
Several bacterial and fungal isolates from a variety of
soils have been shown to degrade the N-methoxy-N-methylurea or N,N-dimethylurea side chains of many phenylurea herbicides [15,19,23,31,46^50]. Nevertheless, until recently there were no reports of de¢ned mixed cultures or
micro-organisms in pure culture able to metabolise the
phenyl structure. It has been suggested that the lack of
success in isolating pure cultures of phenylurea-mineralising bacteria could be related to the involvement of consortia in the degradation processes leading to the mineralisation [26]. This may explain why many studies have been
unsuccessful in isolating phenylurea-degrading bacteria
from mixed cultures readily performing growth-linked
mineralisation of phenylurea compounds [25,28].
Recent studies have indicated that co-operative metabolic activities may be involved in the degradation of
IPU [51], linuron and metobromuron [26,52] by bacterial
consortia derived from previously treated soils. A bacterial
consortium degrading linuron and metobromuron when
provided as the only carbon and nitrogen sources was
enriched and characterised by El-Fantroussi [26]. A consortial degradation involving cross-feeding with aniline
metabolites (3,4-dichloroaniline or 4-bromoaniline) between di¡erent bacteria was strongly indicated by pro¢ling
the mixed bacterial culture using denaturing gradient gel
electrophoresis (DGGE) of polymerase chain reaction
(PCR)-ampli¢ed 16S rRNA genes. Sequencing of DGGE
bands appearing during di¡erent steps in the degradation
revealed proliferation of a strain resembling Variovorax
sp. during metabolism of the aniline metabolites, but attempts to isolate pure cultures able to degrade the herbicides or the aniline derivatives were unsuccessful. Later,
Dejonghe et al. [52] isolated three bacteria from the consortium, identi¢ed as Variovorax paradoxus, Delftia acidovorans and Pseudomonas sp. respectively, which degraded
FEMSEC 1524 2-6-03
6
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
the herbicides when combined, but no details were provided on the role of each member in the consortium.
Srensen et al. [51] described a co-culture performing rapid and extensive growth-linked mineralisation of IPU
when provided as the sole carbon and nitrogen source.
The co-culture consisted of a Sphingomonas sp. (designated strain SRS2) and an unknown soil bacterium (designated strain SRS1) both enriched and isolated from a
British agricultural soil. Sphingomonas sp. SRS2 was the
IPU-mineralising member [29] and it was hypothesised
that strain SRS1 provided amino acids to Sphingomonas
sp. SRS2 and in return grew on cell debris or unknown
metabolites without any direct involvement in the metabolism of IPU [51].
cus, Rhizoctonia solani and Aspergillus niger [31]. Some of
the metabolites created by fungal degradation are more
easily metabolised by soil bacteria than are the herbicides,
and the possibility for a consortial mineralisation involving both fungal and bacterial activity has been suggested
[28]. The role of interactions between bacteria and fungi in
the in situ degradation of phenylurea herbicides remains to
be elucidated.
5. Metabolic pathways for phenylurea herbicide
biodegradation
4.2. Pure cultures of phenylurea-degrading micro-organisms
5.1. General pathways for the metabolism of
N-methoxy-N-methyl- and N,N-dimethyl-substituted
phenylurea herbicides
Several bacterial strains able to partially degrade various phenylurea herbicides have been isolated from soil
[46,49,50,53]. One linuron-degrading isolate was identi¢ed
as a Bacillus sphaericus strain [46] and two diuron-degrading strains as Arthrobacter sp., both closely related to Arthrobacter globiformis [50,53]. Roberts et al. [49] screened
36 bacteria isolated from the Deep Slade agricultural ¢eld
(see Section 6). All were apparently able to degrade IPU.
The strains formed 16 distinct phenotypes, with one group
of isolates identi¢ed as Pseudomonas £uorescens. Later
studies have however shown that each of these cultures
harboured one common additional strain and that it was
this component which performed the actual metabolism of
IPU. This strain was subsequently identi¢ed as a member
of the genus Sphingomonas [27]. These co-cultures greatly
resemble the co-culture described by Srensen et al. [51],
where the mineralisation of IPU is performed by a Sphingomonas sp. supported by a secondary strain not having
any direct involvement in the metabolism.
Sphingomonas sp. strain SRS2, which has recently been
isolated from the Deep Slade ¢eld [29], is the ¢rst described bacterium able to mineralise the phenyl structure
of a phenylurea herbicide. When 14 C-IPU is provided as
the sole source of nitrogen and carbon SRS2 is able to
metabolise the compound to 14 CO2 and biomass. Subsequently a second IPU-mineralising bacterium (designated
strain F35) was isolated from the Deep Slade agricultural
soil [27]. Strain F35 has been identi¢ed as a Sphingomonas
sp. greatly resembling strain SRS2 phylogenetically and in
regard to its metabolism of IPU [27].
Several soil fungi are able to degrade the N-methoxy-Nmethyl- and N,N-dimethyl side chains of di¡erent phenylurea herbicides including IPU, diuron, linuron, monuron,
metobromuron and chlorotoluron (e.g. [19,31,47,48]),
however no fungal mineralisation of the phenyl ring has
so far been reported. The fungi able to partly degrade
phenylurea herbicides cover a broad range of di¡erent
species including Cunninghamella elegans, Mortierella isabellina [19], Talaromyces wortmanii [47], Rhizopus japoni-
The metabolic pathways involved in the degradation of
di¡erent phenylurea herbicides have great similarities, generally involving successive N-demethylation of the N,Ndimethylurea-substituted compounds, and N-demethylation and N-demethoxylation of the N-methoxy-N-methylsubstituted compounds (e.g. [1,24,31,47,54^57]). Following
these stepwise N-dealkylations, the metabolites are hydrolysed to aniline-based metabolites [24,54,56] that subsequently may be further degraded [26,29]. Two A. globiformis strains (designated D47 and N2) and one B. sphaericus
strain (ATCC 12123) isolated from soils are however capable of performing a direct hydrolysis of a broad range
of phenylurea herbicides to their aniline derivatives
[20,58,59] (Fig. 1, pathway II, step 4) thereby by-passing
steps 1^3 presented in Fig. 1, pathway I.
Degradation beyond the aniline-based metabolites is not
well-known for phenylurea herbicides. However, as aniline-based compounds reach the environment from many
other sources, including paints, dyes, plastics and pharmaceutical products, the metabolism of a broad range of
similar compounds by de¢ned micro-organisms and in different environmental samples has been intensively studied
(e.g. [41,60]). Generally aniline compounds are relatively
easily degraded by micro-organisms, however various substitutions to the ring structure, e.g. halogen or nitro
groups, may prevent or delay complete mineralisation.
The general metabolic pathway for metabolism of aniline
involves oxidative deamination to give catechol, which
may be further degraded by di¡erent ring cleavage pathways [41,60]. Under anoxic conditions other processes
such as reductive deamination or dehalogenation may initiate the degradation of aniline metabolites (e.g. [61]). Aniline compounds may also undergo spontaneous chemical
reactions to form various polymerisation products [35,41].
A recent study has shown that 4IA may react with the
humic monomer catechol forming a trimer product identi¢ed as a disubstituted ortho-quinone [35] (Fig. 2, step
14). Another study has shown that 4,4-diisopropylazobenzene, formed from polymerisation of 4IA, may accumulate
FEMSEC 1524 2-6-03
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
in soils during IPU degradation [30,62] (Fig. 2, step 15).
Occurrence of quinone and azo compounds has also been
reported during degradation of chlortoluron [57], and it
seems likely that similar compounds will be produced by
the reactions of other phenylurea-derived aniline metabolites in the soil environment. These polymerisation products probably represent dead-end metabolites, as they have
low biodegradability in di¡erent soil environments [35,62].
5.2. Initial degradation steps involved in the metabolism of
IPU in soils and by isolated micro-organisms
The metabolism of IPU can be initiated by N-demethylation of the N,N-dimethylurea side chain, leading to the
metabolite MDIPU (Fig. 2, step 1). MDIPU has in numerous laboratory and ¢eld studies been reported to be
the metabolite occurring in the highest concentration following IPU treatment of agricultural soils [21,22,30,31,34,
38,62,63]. MDIPU is also the main metabolite occurring
during metabolism of IPU by pure cultures of soil fungi
and bacteria [29,31,49]. Two alternative metabolic pathways involving initial hydroxylation of the isopropyl side
chain, resulting in 1-OH-IPU or 2-OH-IPU (Fig. 2, steps 6
and 7), have also been described in mixed bacterial cultures and in agricultural soils [21,34]. 1-OH-IPU has only
been measured in bacterial cultures derived from soil, but
no e¡ort was made to characterise the degradation mechanism or the organisms involved [34]. 1-OH-IPU has been
reported as a dead-end metabolite without any further
degradation, however its environmental importance remains to be clari¢ed, as this compound has only been
detected in mixed bacterial cultures in a single study
[34]. Detection of both MDIPU and 2-OH-IPU in soil
porewater and surface runo¡ following IPU treatment of
an agricultural ¢eld shows that both pathways can be
active in situ [21]. Low concentrations of the metabolites
DDIPU and 4IA have been detected in IPU-treated agricultural soils [28,34,63] and during the mineralisation of
IPU by Sphingomonas sp. strain SRS2 [29]. A variety of
other hydroxylated metabolites (Fig. 2, steps 8^13) have
been detected during degradation of IPU in di¡erent soils
[21,30,34,63], but a limited amount of information is available on these hydroxylated compounds, and no micro-organism producing these has so far been isolated and
studied in pure culture.
The metabolic pathway proposed for the mineralisation
of IPU by Sphingomonas sp. strain SRS2 (Section 4.2)
involves successive N-demethylations of IPU to MDIPU
and DDIPU, followed by cleavage of the urea side chain
to 4IA (Fig. 2, steps 1^3). 4IA is eventually metabolised to
CO2 and biomass [29]. Previously a mixed bacterial culture
able to mineralise MDIPU and 4IA, but not able to degrade IPU or DDIPU, was enriched from a Danish agricultural soil [28]. A metabolic pathway involving cleavage
of the methylurea group of MDIPU directly to 4IA (Fig.
2, step 5), di¡ering from the N-demethylation to DDIPU
7
performed by Sphingomonas sp. strain SRS2, may thus be
active in the mixed bacterial culture. A. globiformis strain
D47, which is able to degrade the phenylurea herbicides
diuron, linuron, monolinuron, metoxuron and IPU directly to their respective aniline derivatives (Fig. 2, step
4), has been isolated from the Deep Slade agricultural ¢eld
[23,53]. A. globiformis D47 transformed IPU by a single
step involving hydrolytic cleavage of the N,N-dimethylurea side chain to 4IA [59]. A. globiformis strain N2, apparently performing a similar one-step degradation of IPU,
diuron and chlorotoluron to their respective aniline metabolites, has recently been isolated from a French garden
soil that had been treated for several years with diuron
[20,50]. However, none of these strains degraded the phenylurea-derived aniline metabolites produced.
5.3. Genes and enzymes involved in bacterial phenylurea
metabolism
As only a few micro-organisms able to metabolise phenylurea herbicides have been isolated at present, knowledge of the catabolic genes and enzymes involved in phenylurea degradation is lacking. Some studies with
phenylurea-degrading bacteria indicate enzymatic speci¢city related to N-methoxy-N-methyl-substituted phenylurea
herbicides such as linuron and metobromuron [26,58]. An
aryl acylamidase hydrolysing certain phenylurea compounds to their aniline derivatives (Fig. 1, step 4) puri¢ed
from B. sphaericus strain ATCC 12123 had speci¢city related to N-methoxy-N-methyl-substituted phenylurea herbicides, but no activity towards N,N-dimethyl-substituted
phenylurea herbicides such as diuron, monuron and £uometuron [45,58]. El-Fantroussi [26] found a similar speci¢city for degradation of N-methoxy-N-methyl-substituted
herbicides in a recent study of a bacterial consortium enriched from agricultural soil. In contrast to B. sphaericus
ATCC 12123, further degradation of the aniline metabolites was observed with the consortium, thus suggesting
mineralisation of the N-methoxy-N-methyl-substituted
herbicides. Sphingomonas sp. SRS2 apparently has a speci¢city for N,N-dimethylurea-substituted phenylurea herbicides such as IPU, diuron and chlorotoluron, but with no
degradation activity towards the N-methoxy-N-methylsubstituted linuron [29]. The enzymes involved in the
metabolism of the N,N-dimethyl-substituted phenylurea
herbicides by Sphingomonas sp. SRS2 are presently not
studied in detail.
In comparison to the aryl acylamidase from B. sphaericus ATCC 12123, the hydrolase enzyme involved in the
cleavage of phenylurea herbicides to their aniline metabolites by A. globiformis D47 (Section 5.2) has a much
broader substrate speci¢city. The A. globiformis D47 phenylurea hydrolase degrades both N-methoxy-N-methyland N,N-dimethyl-substituted phenylurea herbicides [53,
59]. Recent studies have revealed that the gene encoding
the phenylurea hydrolase, designated gene puhA, is plas-
FEMSEC 1524 2-6-03
8
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
mid-associated in A. globiformis D47 [59]. Gene puhA is
currently the only known gene shown to be associated
with bacterial phenylurea metabolism.
6. Biodegradation of IPU within the Deep Slade
agricultural ¢eld
Recent studies at Deep Slade agricultural ¢eld at HRIWellesbourne, Warwickshire, UK have provided new
understanding of processes underlying spatial variability
in the adaptation of microbial communities to metabolise
phenylurea herbicides (e.g. [22,23,27,29,36,49,59]). The
Deep Slade ¢eld is a research site that has been treated
with IPU regularly for over 20 years. An initial study
reporting accelerated biodegradation of IPU in agricultural soils following successive treatments included soil from
the Deep Slade ¢eld, along with soils from nine other
agricultural ¢elds [22]. This study revealed that IPU was
biodegraded unusually rapidly in the Deep Slade soil,
which stimulated further studies into the underlying microbial processes.
6.1. In-¢eld spatial heterogeneity in the biodegradation of
IPU
Studies have shown that the potential for biodegradation of IPU is unevenly distributed across agricultural
¢elds [36,64^66]. Beck et al. [64] found considerable spatial
heterogeneity in the dissipation rate of IPU in a ¢eld experiment with clay soils. The time required for 50% reduction in the IPU concentration ranged from 31 to 483 days
in 25 di¡erent sampling areas, each 1 m2 , within an agricultural plot of 40U28 m. Walker et al. [65] reported a
substantial spatial variability in the degradation rate of
IPU within the Deep Slade ¢eld by comparing 30 di¡erent
locations, each 500 g, from a grid area of 200U250 m.
Deep Slade ¢eld possessed sandy loam with apparently
uniform soil characteristics [65]. However, shorter IPU
half-lives were positively correlated with increasing soil
pH (ranging from 6.1 to 7.6) and soil microbial biomass
[65]. Further studies by Bending et al. [36] focussed on
areas within Deep Slade ¢eld with known contrasting
IPU degradation rates. Fast degradation of IPU was associated with rapid mineralisation of 14 C-IPU to 14 CO2
and with incorporation of a higher amount of 14 C into
microbial biomass compared to the slowly degrading
soils [36]. All soils harboured IPU-metabolising microorganisms prior to IPU application, but only the soils
with fast degradation were associated with a proliferation of degrading organisms during IPU metabolism [27,
36].
A signi¢cant negative relationship between soil pH (6.1^
7.5) and the degradation rate of IPU, similar to the one
presented from the Deep Slade studies, was recently reported within another British agricultural ¢eld [66]. This
study also revealed an in-¢eld spatial heterogeneity in the
distribution of IPU and chlorotoluron degradation potential comparable to the previous reports based on the Deep
Slade studies. These similarities may suggest that the in¢eld degradation patterns reported from Deep Slade may
be relevant for a broader range of agricultural ¢elds, and
that the microbial degradation potential towards phenylurea herbicides can be linked with key soil parameters
related to pH.
6.2. Linking pure culture studies with in-¢eld spatial
variability in the biodegradation of IPU
Several of the phenylurea-degrading bacteria presented
so far (Section 4.2) have been enriched and isolated from
the Deep Slade soil [23,27,29,49]. There is a general lack of
knowledge on the role of degradative isolates in in situ
degradation of xenobiotics in soils. However, studies involving Deep Slade ¢eld have provided new insights into
the distribution and activities of phenylurea-degrading micro-organisms in situ.
Bending et al. [27] used DGGE analysis of PCR-ampli¢ed 16S rRNA genes from soil-extracted DNA to study
bacterial community dynamics during IPU degradation in
soil from Deep Slade ¢eld. In the soils showing rapid IPU
degradation, proliferation of speci¢c IPU-metabolising
populations was indicated by most-probable-number analysis and by the appearance of two DGGE bands during
IPU degradation. Sequencing and alignment revealed that
one of the DGGE bands had a 100% similarity with the
partial 16S rRNA gene sequence of IPU-metabolising
Sphingomonas sp. strain SRS2 (Section 4.2) in a 130-bp
overlap. This indicates that strain SRS2, or closely related
Sphingomonas strains, is involved in the fast degradation
of IPU observed in soils from the Deep Slade ¢eld. Sequence analysis of the second DGGE band suggested that
this represents an unknown Sphingomonas sp. [27]. Another IPU-metabolising Sphingomonas sp. (strain F35; Section 4.1) has recently been isolated from the Deep Slade
soil [27], but none of the known phenylurea-degrading
strains has so far been assigned to the second DGGE
band. In the soils with slow degradation of IPU, no consistent changes in the DGGE banding pro¢les were observed [27].
In pure culture the speci¢c mineralisation rate of 14 CIPU to 14 CO2 by strains SRS2 and F35 is a¡ected by the
pH of the liquid medium, with signi¢cantly faster rates at
increasing pH in the range 6.5^7.5 [27]. Similar results
were obtained in inoculation experiments when strain
SRS2 was introduced into two agricultural soils having a
natural pH of 6.5 and 7.5, respectively. Collectively these
results suggest that the bacteria involved in the IPU degradation across the Deep Slade ¢eld are a¡ected by the
spatial heterogeneity in the soil pH, which thereby determines the in situ degradation rate of IPU, and its susceptibility to leaching losses to groundwater.
FEMSEC 1524 2-6-03
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
7. Summary and conclusions
Studies on the environmental fate of phenylurea herbicides have attracted increasing attention in recent years
due to rising concern about the impact of water contamination on environmental and public health. The degradation of phenylurea herbicides following application to
agricultural ¢elds is predominantly microbial, and evidence suggests that surface soils act as reactive microbial
barriers determining the amounts and types of herbicide
residues transported from the soil by water movement.
This active biodegradation zone may be the only protection against contamination of groundwater with phenylurea herbicides and their metabolites, as recent research
implies that the potential for degradation decreases signi¢cantly below the surface soil, and so far no evidence for
extensive biodegradation in subsurface environments has
been presented. This review has provided an overview of
the current knowledge on the microbial processes determining the extent and rate of degradation of phenylurea
herbicides in agricultural soils.
Recent studies suggest that enhanced biodegradation of
IPU in agricultural soils may develop with successive ¢eld
treatments. However, considerable in-¢eld spatial heterogeneity in the distribution of the IPU-metabolising community can still exist. Detailed studies including culturedependent and molecular approaches with soils from the
Deep Slade ¢eld have provided understanding of the microbial mechanisms underlying this spatial variability.
Areas showing rapid degradation of IPU had relatively
high pH and IPU degradation was associated with proliferation of a small group of closely related IPU-degrading
Sphingomonas spp. Areas with slow IPU biodegradation
had relatively low pH and either slow or no proliferation
of IPU-degrading micro-organisms. This pattern was successfully linked to pure culture studies with isolated IPUdegrading Sphingomonas spp. strains. These results suggest
that even within agricultural ¢elds which have received
extensive applications of phenylurea herbicides, and in
which there has been adaptation of phenylurea-metabolising micro-organisms, areas with unfavourable growth conditions exist. This has implications for the patterns of
leaching of herbicide residues to underlying groundwater
aquifers.
Substantial in-¢eld heterogeneity in degradation potential may only occur for recalcitrant herbicides such as
phenylureas and triazines, where the ability to degrade
the compounds most likely is restricted to small groups
of micro-organisms. In contrast, more easily biodegradable pesticides that a broad range of micro-organisms
are able to metabolise such as the phenoxyalkanoic acid
herbicides and the carbamate insecticides, will probably
result in a more even in-¢eld distribution of the degradation potential.
It remains to be elucidated whether the correlation between soil pH, presence of speci¢c Sphingomonas spp.
9
strains and IPU degradation is relevant for other IPUtreated agricultural ¢elds. Isolation and characterisation
of more phenylurea-degrading micro-organisms from agricultural soils covering a broader range of geographical
areas may shed some light on the identity, distribution
and degree of divergence among degrading soil micro-organisms and eventually catabolic genes and enzymes.
Acknowledgements
This work was in part supported by the Danish Environmental Research Program (SMP96) - Pesticides and
Groundwater (Denmark), the Department of Environment, Food and Rural A¡airs (UK), and the Biotechnology and Biological Sciences Research Council (UK). We
thank Dr P.H. Pritchard (National Environmental Research Institute of Denmark, Department of Microbial
Ecology and Biotechnology, Roskilde, Denmark) for encouraging the production of this review. Dr K. Johnsen
(Novozymes, Kalundborg, Denmark) and Ms P. Simpson
are thanked for fruitful discussions and constructive comments on the manuscript.
References
[1] Field, J.A., Reed, R.L., Sawyer, T.E. and Martinez, M. (1997) Diuron and its metabolites in surface water and ground water by solid
phase extraction and in-vial elution. J. Agric. Food Chem. 45, 3897^
3902.
[2] Spliid, H.S. and Kppen, B. (1998) Occurrence of pesticides in Danish shallow groundwater. Chemosphere 37, 1307^1316.
[3] Stangroom, S.J., Collins, C.D. and Lester, J.N. (1998) Sources of
organic micropollutants to lowland rivers. Environ. Technol. 19,
643^666.
[4] Thurman, E.M., Bastian, K.C. and Mollhagen, T. (2000) Occurrence
of cotton herbicides and insecticides in Playa Lakes of the High
Plains of West Texas. Sci. Total Environ. 248, 189^200.
[5] Gerecke, A.C., Canonica, S., Mu«ller, S.R., Schaerer, M. and Schwarzenbach, R.P. (2001) Quanti¢cation of dissolved natural organic matter (DOM) mediated phototransformation of phenylurea herbicides
in lakes. Environ. Sci. Technol. 35, 3915^3923.
[6] Gerecke, A.C., Tixier, C., Bartels, T., Schwarzenbach, R.P. and Mu«ller, S.R. (2001) Determination of phenylurea herbicides in natural
waters at concentrations below 1 ng l31 using solid-phase extraction,
derivatization, and solid-phase microextraction gas chromatographymass spectrometry. J. Chromatogr. A 930, 9^19.
[7] Johnson, A.C., Besien, T.J., Bhardwaj, C.L., Dixon, A., Gooddy,
D.C., Haria, A.H. and White, C. (2001) Penetration of herbicides
to groundwater in an uncon¢ned chalk aquifer following normal
soil applications. J. Contam. Hydrol. 53, 101^117.
[8] Gooddy, D.C., Chilton, P.J. and Harrison, I. (2002) A ¢eld study to
assess the degradation and transport of diuron and its metabolites in
a calcareous soil. Sci. Total Environ. 297, 67^83.
[9] Alexander, M. (1981) Biodegradation of chemicals of environmental
concern. Science 211, 132^138.
[10] Johnson, A.C., Hughes, C.D., Williams, R.J. and Chilton, P.J. (1998)
Potential for aerobic isoproturon biodegradation and sorption in the
unsaturated and saturated zones of a chalk aquifer. J. Cont. Hydrol.
30, 281^297.
FEMSEC 1524 2-6-03
10
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
[11] Larsen, L., Srensen, S.R. and Aamand, J. (2000) Mecoprop, isoproturon, and atrazine in and above a sandy aquifer : vertical distribution of mineralization potential. Environ. Sci. Technol. 34, 2426^
2430.
[12] Kristensen, G.B., Srensen, S.R. and Aamand, J. (2001) Mineralization of 2,4-D, mecoprop, isoproturon and terbutylazine in a chalk
aquifer. Pest Manag. Sci. 57, 531^536.
[13] Environment Agency (2001) Pesticides 2000: a summary of monitoring of the aquatic environment in England and Wales. Environmental
Agency. www.environment-agency.gov.uk.
[14] Environment Agency (2000) Monitoring of pesticides in the environment. Environmental Agency. www.environment-agency.gov.uk.
[15] Hill, G.D., McGahen, J.W., Baker, H.M., Finnerty, D.W. and Bingeman, C.W. (1955) The fate of substituted urea herbicides in agricultural soils. Agron. J. 47, 93^104.
[16] Salvestrini, S., Di Cerbo, P. and Capasso, S. (2002) Kinetics of the
chemical degradation of diuron. Chemosphere 48, 69^73.
[17] Kulshrestha, G. and Mukerjee, S.K. (1986) The photochemical decomposition of the herbicide isoproturon. Pestic. Sci. 17, 489^494.
[18] Remde, A. and Traunspurger, W. (1994) A method to assess the
toxicity of pollutants on anaerobic microbial degradation activity in
sediments. Environ. Toxicol. Water Qual. 9, 293^298.
[19] Tixier, C., Bogaerts, P., Sancelme, M., Bonnemoy, F., Twagilimana,
T., Cuer, A., Bohatier, J. and Veschambre, H. (2000) Fungal biodegradation of a phenylurea herbicide, diuron: structure and toxicity of
metabolites. Pest Manag. Sci. 56, 455^462.
[20] Tixier, C., Sancelme, M., A|«t-A|«ssa, S., Widehem, P., Bonnemoy, F.,
Cuer, A., Tru¡aut, N. and Veschambre, H. (2002) Biotransformation
of phenylurea herbicides by a soil bacterial strain Arthrobacter sp. N2
structure, ecotoxicity and fate of diuron metabolite with soil fungi.
Chemosphere 46, 519^526.
[21] Schuelein, J., Glaessgen, W.E., Hertkorn, N., Schroeder, P., Sandermann, H.Jr. and Kettrup, A. (1996) Detection and identi¢cation of
the herbicide isoproturon and its metabolites in ¢eld samples after a
heavy rainfall event. Int. J. Environ. Anal. Chem. 65, 193^202.
[22] Cox, L., Walker, A. and Welch, S.J. (1996) Evidence for the accelerated degradation of isoproturon in soils. Pestic. Sci. 48, 253^260.
[23] Cullington, J.E. and Walker, A. (1999) Rapid biodegradation of
diuron and other phenylurea herbicides by a soil bacterium. Soil
Biol. Biochem. 31, 677^686.
[24] Bozarth, G.A. and Funderburk Jr., H.H. (1971) Degradation of £uometuron in sandy loam soil. Weed Sci. 19, 691^695.
[25] Roberts, S.J., Walker, A., Parekh, N.R., Welch, S.J. and Waddington, M.J. (1993) Studies on a mixed bacterial culture from soil which
degrades the herbicide linuron. Pestic. Sci. 39, 71^78.
[26] El-Fantroussi, S. (2000) Enrichment and molecular characterization
of a bacterial culture that degrades methoxy-methyl urea herbicides
and their aniline derivatives. Appl. Environ. Microbiol. 66, 5110^
5115.
[27] Bending, G.D., Lincoln, S.D., Srensen, S.R., Morgan, J.A.W., Aamand, J. and Walker, A. (2003) In-¢eld spatial variability in the
degradation of the phenyl-urea herbicide isoproturon is the result
of interactions between degradative Sphingomonas spp. and soil pH.
Appl. Environ. Microbiol. 69, 827^834.
[28] Srensen, S.R. and Aamand, J. (2001) Biodegradation of the phenylurea herbicide isoproturon and its metabolites in agricultural soils.
Biodegradation 12, 69^77.
[29] Srensen, S.R., Ronen, Z. and Aamand, J. (2001) Isolation from
agricultural soil and characterization of a Sphingomonas sp. able to
mineralize the phenylurea herbicide isoproturon. Appl. Environ. Microbiol. 67, 5403^5409.
[30] Pieuchot, M., Perrin-Ganier, C., Portal, J.-M. and Schiavon, M.
(1996) Study on the mineralization and degradation of isoproturon
in three soils. Chemosphere 33, 467^478.
[31] Berger, B.M. (1999) Factors in£uencing transformation rates and
formation of products of phenylurea herbicides in soil. J. Agric.
Food Chem. 47, 3389^3396.
[32] Zablotowicz, R.M., Locke, M.A., Gaston, L.A. and Bryson, C.T.
(2000) Interactions of tillage and soil depth on £uometuron degradation in a Dundee silt loam soil. Soil Till. Res. 57, 61^68.
[33] Kubiak, R., Ellssel, H., Lambert, M. and Eichhorn, K.W. (1995)
Degradation of isoproturon in soil in relation to changes of microbial
biomass and activity in small-scale laboratory and outdoors studies.
Int. J. Anal. Chem. 59, 123^132.
[34] Lehr, S., Gla«ssgen, W.E., Sandermann, H., Beese, Jr.F. and Scheunert, I. (1996) Metabolism of isoproturon in soils originating from
di¡erent agricultural management systems and in cultures of isolated
soil bacteria. Int. J. Environ. Anal. Chem. 65, 231^243.
[35] Scheunert, I. and Reuter, S. (2000) Formation and release of residues
of the 14 C-labelled herbicide isoproturon and its metabolites bound in
model polymers and in soil. Environ. Pollut. 108, 61^68.
[36] Bending, G.D., Shaw, E. and Walker, A. (2001) Spatial heterogeneity
in the metabolism and dynamics of isoproturon degrading microbial
communities in soil. Biol. Fertil. Soils 33, 484^489.
[37] Srensen, S R. and Aamand, J. (2003) Rapid mineralisation of the
herbicide isoproturon in soil from a previously treated Danish agricultural ¢eld. Pest Manag. Sci. (in press).
[38] Gaillardon, P. and Sabar, M. (1994) Changes in the concentrations of
isoproturon and its degradation products in soil and soil solution
during incubation at two temperatures. Weed Res. 34, 243^250.
[39] Bollag, J.-M., Blattmann, P. and Laanio, T. (1978) Adsorption and
transformation of four substituted anilines in soil. J. Agric. Food
Chem. 26, 1302^1306.
[40] Johannesen, H., Srensen, S.R. and Aamand, J. (2003) Mineralization of soil-aged isoproturon and isoproturon metabolites by Sphingomonas sp. strain SRS2. J. Environ. Qual. (in press).
[41] Parris, G.E. (1980) Environmental and metabolic transformation of
primary aromatic amines and related compounds. Residue Rev. 76,
1^30.
[42] Cox, L. and Walker, A. (1999) Studies of time-dependent sorption of
linuron and isoproturon in soils. Chemosphere 38, 2707^2718.
[43] Johnson, A.C., White, C. and Bhardwaj, C.L. (2000) Potential for
isoproturon, atrazine and mecoprop to be degraded within a chalk
aquifer system. J. Cont. Hydrol. 44, 1^18.
[44] Larsen, L. and Aamand, J. (2001) Degradation of herbicides in two
sandy aquifers under di¡erent redox conditions. Chemosphere 44,
231^236.
[45] Attaway, H.H., Paynter, M.J.B. and Camper, N.D. (1982) Degradation of selected phenylurea herbicides by anaerobic pond sediment.
J. Environ. Sci. Health B17, 683^699.
[46] Wallno«fer, P. (1969) The decomposition of urea herbicides by Bacillus sphaericus, isolated from soil. Weed Res. 9, 333^339.
[47] Tweedy, B.G., Loeppky, C. and Ross, J.A. (1970) Metabolism of
3-(p-bromophenyl)-1-methoxy-1-methylurea (metobromuron) by selected soil microorganisms. J. Agric. Food Chem. 18, 851^853.
[48] Vroumsia, T., Steiman, R., Seigle-Murandi, F., Benoit-Guyod, J.-L.
and Khadrani, A. (1996) Biodegradation of three substituted phenylurea herbicides (chlortoluron, diuron, and isoproturon) by soil fungi.
A comparative study. Chemosphere 33, 2045^2056.
[49] Roberts, S.J., Walker, A., Cox, L. and Welch, S.J. (1998) Isolation of
isoproturon-degrading bacteria from treated soil via three di¡erent
routes. J. Appl. Microbiol. 85, 309^316.
[50] Widehem, P., A|«t-A|«ssa, S., Tixier, C., Sancelme, S., Veschambre, H.
and Tru¡aut, N. (2002) Isolation, characterization and diuron transformation capacities of a bacterial strain Arthrobacter sp. N2. Chemosphere 46, 527^534.
[51] Srensen, S.R., Ronen, Z. and Aamand, J. (2002) Growth in coculture stimulates metabolism of phenylurea herbicides isoproturon
by Sphingomonas sp. SRS2. Appl. Environ. Microbiol. 68, 3478^
3485.
[52] Dejonghe, W., Goris, J., El-Fantroussi, S., De Vos, P., Verstraete, W.
and Top, E.M. (2000) Microbiological and genetic characterization of
a linuron degrading consortium. Med. Fac. Landb. Univ. Gent 65,
69^73.
FEMSEC 1524 2-6-03
S.R. S-rensen et al. / FEMS Microbiology Ecology 45 (2003) 1^11
[53] Turnbull, G.A., Cullington, J.E., Walker, A. and Morgan, J.A.W.
(2001) Identi¢cation and characterization of a diuron-degrading bacterium. Biol. Fertil. Soils 33, 472^476.
[54] Dalton, R.L., Evans, A.W. and Rhodes, R.C. (1966) Disappearance
of diuron from cotton ¢eld soils. Weeds 14, 31^33.
[55] Ross, J.A. and Tweedy, B.G. (1973) Degradation of four phenylurea
herbicides by mixed populations of microorganisms from two soil
types. Soil Biol. Biochem. 5, 739^746.
[56] Geissbu«hler, H., Martin, H. and Voss, G. (1975) The substituted
ureas. In: Herbicides ^ Chemistry, Degradation and Mode of Action,
2nd edn. (Kearney, P.C. and Kaufman, D.D., Eds.), pp. 209^291.
Marcel Dekker, New York.
[57] Smith, A.E. and Briggs, G.G. (1978) The fate of the herbicide chlortoluron and its possible degradation products in soils. Weed Res. 18,
1^7.
[58] Engelhardt, G., Wallno«fer, P.R. and Plapp, R. (1973) Puri¢cation
and properties of an aryl acylamidase of Bacillus sphaericus, catalyzing the hydrolysis of various phenylamide herbicides and fungicides.
Appl. Microbiol. 26, 709^718.
[59] Turnbull, G.A., Ousley, M., Walker, A., Shaw, E. and Morgan,
J.A.W. (2001) Degradation of substituted phenylurea herbicides by
Arthrobacter globiformis strain D47 and characterization of a plasmid-associated hydrolase gene, puhA. Appl. Environ. Microbiol. 67,
2270^2275.
11
[60] Lyons, C.D., Katz, S. and Bartha, R. (1984) Mechanisms and pathways of aniline elimination from aquatic environments. Appl. Environ. Microbiol. 48, 491^496.
[61] Travkin, V., Baskunov, B.P., Golovlev, E.L., Boersma, M.G., Boeren, S., Vervoort, J., van Berkel, W.J.H., Rietjens, I.M.C.M. and
Golovleva, L.A. (2002) Reductive deamination as a new step in the
anaerobic microbial degradation of halogenated anilines. FEMS Microbiol. Lett. 209, 307^312.
[62] Perrin-Ganier, C., Schiavon, F., Morel, J.-L. and Schiavon, M.
(2001) E¡ect of sludge-amendment or nutrient addition on the biodegradation of the herbicide isoproturon in soil. Chemosphere 44,
887^892.
[63] Mudd, P.J., Hance, R.J. and Wright, S.J.L. (1983) The persistence
and metabolism of isoproturon in soil. Weed Res. 23, 239^247.
[64] Beck, A.J., Harris, G.L., Howse, K.R., Johnston, A.E. and Jones,
K.C. (1996) Spatial and temporal variation of isoproturon residues
and associated sorption/desorption parameters at the ¢eld scale. Chemosphere 33, 1283^1295.
[65] Walker, A., Jurado-Exposito, M., Bending, G.D. and Smith, V.J.R.
(2001) Spatial variability in the degradation rate of isoproturon in
soil. Environ. Pollut. 111, 407^415.
[66] Walker, A., Bromilow, R.H., Nicholls, P.H., Evans, A.A. and Smith,
V.J.R. (2002) Spatial variability in the degradation rates of isoproturon and chlorotoluron in a clay soil. Weed Res. 42, 39^44.
FEMSEC 1524 2-6-03