C H A P T E R
F I V E
Ecology of Denitrifying Prokaryotes
in Agricultural Soil
Laurent Philippot,* Sara Hallin,† and Michael Schloter‡
Contents
1. Introduction
2. Agronomical and Environmental Importance of Denitrification
2.1. Consequences of denitrification for agriculture
2.2. Impact of denitrification on the environment and
human health
3. Who are the Denitrifiers?
3.1. Denitrifiers and nitrate reducers
3.2. Denitrifying populations
4. Assessing Denitrifiers Density, Diversity, and Activity
4.1. Measuring denitrification and N2O emissions
4.2. Resolving diversity of denitrifiers
4.3. Quantification of denitrifiers
5. Natural Factors Causing Variations in Denitrification
5.1. Temperature and water
5.2. Freeze–thaw cycles
5.3. Dry–wet cycles
6. Denitrification in the Rhizosphere of Crops
6.1. Crops as a factor influencing denitrifiers
6.2. Impact of crop species, crop cultivars, and transgenic plants
7. Impact of Fertilization on Denitrification
7.1. Fertilization affects denitrification
8. Effect of Environmental Pollution on Denitrifiers
8.1. Pollution affects denitrification
8.2. Pesticides
8.3. Heavy metals
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INRA, University of Burgundy, Soil and Environmental Microbiology, Dijon, France
Department of Microbiology, Swedish University of Agricultural Sciences, Uppsala, Sweden
GSF-National Research Center for Environment and Health, Institute for Soil Ecology, Oberscheissheim,
Germany
Advances in Agronomy, Volume 96
ISSN 0065-2113, DOI: 10.1016/S0065-2113(07)96003-4
#
2007 Elsevier Inc.
All rights reserved.
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Laurent Philippot et al.
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9. Conclusions and Outlook
References
Denitrification is a microbial respiratory process during which soluble nitrogen
oxides are used as an alternative electron acceptor when oxygen is limiting. It
results in considerable loss of nitrogen, which is the most limiting nutrient for
crop production in agriculture. Denitrification is also of environmental concern,
since it is the main biological process responsible for emissions of nitrous
oxide, one of the six greenhouse gases considered by the Kyoto protocol. In
addition to natural variations, agroecosystems are characterized by the use of
numerous practices, such as fertilization and pesticide application, which can
influence denitrification rates. This has been widely documented in the literature, illustrating the complexity of the underlying mechanisms regulating
this process. In the last decade, however, application of molecular biology
approaches has given the opportunity to look behind denitrification rates and
to describe genes, transcripts, and enzymes responsible for the process. In
order to reduce denitrification in arable soil, it is important to understand how
different factors influence denitrification and how the denitrifier community
structure is related to in situ activity. This chapter focuses on the impact of
natural events as well as agricultural practices on denitrifying microorganisms.
1. Introduction
In nature, nitrogen is present in different oxidation forms ranging from
reduced compounds, for example, –3 in ammonia, to fully oxidized, for
example, þ5 in nitrate (NO
3 ).
The conversion between these different forms of nitrogen is mainly
mediated by microorganisms (Fig. 1). The major pool of nitrogen is found
Nitrogen fixation
DNRA
NH+
4
NO2
Nitrification
NO3
NH3
NO2
NH2OH
N2
NO
N2O
N2
Denitrification
NH2OH
Anammox NH+
4
N2H2
N2
Figure 1 Microbial processes contributing to the biological nitrogen cycle.
Soil Denitrifiers
251
in the atmosphere as dinitrogen gas. It can be converted into ammonia by
symbiotic as well as free-living prokaryotes (Bacteria and Archeae) called
diazotrophs, which can break the triple covalent bond of dinitrogen gas. This
process is named biological nitrogen fixation. Ammonia itself can be oxidized
into NO
3 during a two-step process called nitrification. The NO3
produced may be reduced either to dinitrogen gas via denitrification or by
dissimilatory NO
3 reduction to ammonium (DNRA). These steps form the
major parts of the inorganic nitrogen cycle in soils. Other reactions, like the
anaerobic ammonia oxidation (Anammox), where nitrite (NO
2 ) is reduced
to dinitrogen gas using ammonia as an inorganic electron donor (Mulder
et al., 1995), have been shown to occur in several environments. Nevertheless, it has not been proven yet, that Anammox plays a major role in soil
ecosystems (Jetten, 2001). Ammonia and NO
3 can be used by most living
cells to produce organic forms of nitrogen, like proteins, amino acids, and so
on, which are essential for life. During decay of biomass (plants, animals,
fungi, bacteria), these organic nitrogen forms are degraded and transferred
into ammonia again. Therefore, ammonia is the link between organic and
inorganic nitrogen cycle.
Together these processes form the global nitrogen cycle and microorganisms are essential for maintaining the balance between reduced and oxidized
forms of nitrogen. In many soil ecosystems, nitrogen is often the limiting
nutrient for plant growth and it is continuously lost by denitrification, soil
erosion, leaching, and ammonia volatilization. Nitrogen losses through
ammonia volatilization and denitrification are significant factors to consider
when developing nitrogen management strategies in agricultural cropping
systems. In particular, denitrification leads to nitrogen loss from soil, and results
in the release of nitrous oxide (N2O), which is among the six greenhouse
gases considered by the Kyoto protocol on climate change in 1997. Thus,
increasing our knowledge of microbial communities involved in the nitrogen
cycle is important, not only for increasing plant available nitrogen, but also for
reducing the negative impact of agriculture on the environment.
Denitrification can be defined as a microbial respiratory process during
which soluble nitrogen oxides are used as alternative electron acceptor
when oxygen is not available for aerobic respiration. It consists in the
sequential reduction of NO
3 into dinitrogen in four steps concomitant
with energy conservation (Fig. 2). This reduction of NO
3 by bacteria was
discovered in the second-half of the nineteenth century by Gayon and
Dupetit (1886). Substantial progress has been made during the last
20 years concerning the biochemistry and genetic of denitrification, which
has been summarized in a number of comprehensive reviews (Berks et al.,
1995; Philippot, 2002a; Zumft, 1997). Briefly, two types of molybdoen
zymes catalyzing the first step of the pathway, the reduction of NO
3 to NO2
have been described: a membrane bound (Nar) and a periplasmic (Nap)
NO
3 reductases. Both types of enzymes can be present in the same strain
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Laurent Philippot et al.
N2
- Nitrous oxide reductase reductase
(nosZ )
- Nitric oxide reductase
(norB)
- Quinol nitric oxide reductase
(qnorB)
- Cd1 nitrite reductase
(nirS)
- Cu nitrite reductase
(nirK )
N2O
NO
NO2−
- Membrane bound nitrate reductase (narG)
- Periplasmic nitrate reductase
(napA)
NO3−
Figure 2 The denitrification cascade with the different reductases and name of the
genes encoding the corresponding catalytic subunits (in parentheses).
(Carter et al., 1995; Roussel-Delif et al., 2005). The reduction of soluble
NO
2 into gaseous nitric oxide (NO), the key step in the denitrification
cascade, can be catalyzed by evolutionary unrelated enzymes that are different
in terms of structure and of prosthetic metals—a copper (NirK) and a cyto
chrome cd1 (NirS) NO
2 reductase. In contrast to the NO3 reductases, bacteria
carry either the copper or the cd1 NO2 reductase but the two enzymes are
functionally equivalent (Glockner et al., 1993). Reduction of NO into nitrous
oxide is also catalyzed by two types of enzymes: one NO reductase receives the
electrons from cytochrome c or pseudoazurin (cNor) and the other from a
quinol pool (qNor). The last step of the denitrification cascade, reduction of
N2O into dinitrogen gas, is performed by the multicopper homodimeric N2O
reductase (NosZ), which is located in the periplasm in Gram-negative
bacteria.
Soil Denitrifiers
253
The general requirements for biological denitrification are: (1) the
presence of bacteria possessing the metabolic capacity; (2) suitable electron
donors such as organic carbon compounds; (3) anaerobic conditions or
restricted O2 availability; and (4) presence of N-oxides (NO
3 , NO2 ,
NO, or N2O) as terminal electron acceptors. The process of denitrification
is therefore generally promoted under anaerobic conditions, high levels of
soil NO
3 , and a readily available source of carbon.
In this chapter, we will highlight the agronomical and environmental
importance of denitrification and give a brief overview of the methods used
to assess denitrifier activity, diversity, and density. The activity and diversity
of denitrifiers is discussed in relation to natural factors, plant effects in crop
production, fertilization regimes, or use of pesticides.
2. Agronomical and Environmental Importance
of Denitrification
2.1. Consequences of denitrification for agriculture
Denitrification leads to considerable nitrogen losses in agriculture. The
losses tend to increase with fertilization, and between 0% and 25% of the
applied nitrogen can end up as nitrogen gas or N2O, thus limiting crop
production (Aulakh et al., 1992; De Klein and Van Logtestijn, 1994; Mogge
et al., 1999). Studies have shown that up to 340 kg N ha1 can be lost
through denitrification during 1 year under extreme conditions, although
values in the range 0–200 kg N ha1 year1 are more normal (Hofstra and
Bouwman, 2005). The values obtained depend highly on the methods used
to determine denitrification rates (Section 4.1). Models have estimated the
total annual denitrification for the global agricultural area (excluding leguminous crops) to be 22–87 Tg nitrogen (Drecht et al., 2003; Hofstra and
Bouwman, 2005).
Intensively cultivated soils have higher denitrification activity compared
with native noncultivated soils. Nevertheless, denitrification events in the
field occur irregularly in time and space because of weather conditions,
heterogeneity of soil conditions, and management practices. The highest
rates are often measured in spring and fall, which indicates that soil water
status is a strong controlling factor. Hence, flood-irrigated cropping systems
are especially prone to denitrification and recovery of fertilized nitrogen is
often poor (Aulakh et al., 2001; Mahmood et al., 2000, 2005). To minimize
the nitrogen losses, the feasible option is to focus on agricultural practices.
After compiling 336 datasets on denitrification measurements, Hofstra and
Bouwman (2005) demonstrated that crop-type, fertilizer-type, and nitrogen
application rate were the most significant management-related factors influencing denitrification in agricultural soils. These factors not only affect the
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Laurent Philippot et al.
nitrogen availability and the form of available nitrogen in soil, but also affect
the type and amount of carbon available for denitrification.
2.2. Impact of denitrification on the environment and
human health
Denitrification together with nitrification are considered as the primary
biological sources of N2O, which exhibits a global warming potential
300 times higher than that of carbon dioxide as defined by the Intergovernmental Panel on Climate Change (IPCC) and contributes up to 6% of
the anthropogenic greenhouse effect (Cicerone, 1989). N2O also participates in depletion of the stratospheric ozone layer through stratospheric NO
production (Tabazadeh et al., 2000; Waibel et al., 1999). N2O emission by
denitrification is the net result of the balance between production and
reduction of N2O by denitrifying bacteria.
Soil ecosystems are the dominant sources of atmospheric N2O (Conrad,
1996), contributing to 70% (10 Tg year1) of the total annual global emission with about 6.3 Tg year1 from agricultural soils, animal production, and
other agricultural activities (Mosier et al., 1998). From the preindustrial
period to our days, the atmospheric concentration of N2O increased from
0.275 to 0.314 ppm with an actual increase rate of 0.3% per year. This has
been attributed to the increased use of nitrogen fertilizers (Skiba and Smith,
2000). Only between 1960 and 1995, there was a sevenfold increase in
fertilization (Tilman et al., 2002). The 1996 IPCC guidelines used a fixed
N2O emission rate of 1.25% for all nitrogen applied as fertilizer (Houghton
et al., 1996). However, studies suggested N2O emissions from agricultural
soils might be twice as high as IPCC estimates (Giles, 2005).
Denitrification is also of interest for nitrogen removal in agricultural
drainage and runoff water, groundwater, wastewater, and drinking water,
the latter being of a special concern for human health. The removal of
nitrogen in the form of ammonia and NO
3 is effected through the
biological oxidation of nitrogen from ammonia (nitrification) to NO
3,
followed by denitrification. Nitrogen gas is then released to the atmosphere
and thus removed from the water. High NO
3 concentrations in drinking
water are toxic, especially to infants under 6 months. However, NO
3 itself
does not normally cause health problems unless it is reduced to NO
2 by
bacteria that live in the digestive tract. As NO
2 enters the blood stream, it
reacts with hemoglobin to form methemoglobin, and oxygen transportation
is blocked. This causes asphyxiation, a disease commonly called ‘‘blue baby
syndrome’’ or methemoglobinemia. Nitrate in groundwater originates primarily from fertilizers, septic systems, and manure storage or application.
Thus, fertilizer nitrogen that is not taken up by plants, volatilized, denitrified, or carried away by surface run-off leaches to the groundwater in the
form of NO
3 . The World Heath Organization has stipulated a safe upper
limit of 45 mg NO3 liter1 in drinking water for human consumption.
Soil Denitrifiers
255
3. Who are the Denitrifiers?
3.1. Denitrifiers and nitrate reducers
Many soil prokaryotes can denitrify and exhibit a variety of reduction pathways for nitrogenous oxides. Both cultivation-dependent and -independent
methods showed that the proportion of denitrifiers represent up to 5% of
the total soil microbial community (Henry et al., 2004, 2006; Tiedje, 1988),
thus outranking other functional groups involved in the N-cycle such as
diazotrophs or nitrifiers. Some microorganisms produce only nitrogen
gas as end denitrification product, while others give a mixture of N2O
and nitrogen gas, and some only N2O (Stouthamer, 1988). In addition, a
few microorganisms cannot reduce NO
3 and use NO2 as the first electron
acceptor in the denitrification cascade. By contrast, some NO
3 -reducing
bacteria reduce the produced NO
into
ammonium
and
not
into
NO. The
2
dissimilatory NO
reduction
into
ammonium
should
be
distinguished
3
from denitrification, even though it may produce nitrogenous gases as byproducts. Therefore, many NO
3 -respiring ammonium-producing isolates
have been misidentified as denitrifiers. Accordingly, different criteria have
been proposed to identify ‘‘true’’ denitrifiers and to distinguish them from
the NO
3 -respiring, ammonium-producing bacteria (Mahne and Tiedje,
1995): (1) N2O and/or nitrogen gas must be the major end product of NO
3
or NO
2 reduction; and (2) this reduction must be coupled to an increased
in growth yield increase that is greater than when NO
3 or NO2 simply
served as an electron sink. Using these criteria, it is also possible to distinguish bacteria possessing only the NO reductase as a protection against
exogenous or endogenous nitrosative stress (Philippot, 2005).
3.2. Denitrifying populations
More than 60 genera of denitrifying microorganisms have been identified
including archeae and fungi (Table 1). Consequently, the distribution of the
denitrification trait among microorganisms cannot be predicted simply by
the taxonomical affiliation. In addition, while distantly related microorganisms can denitrify, closely related strains can exhibit different respiratory
pathways. For example, analysis of the ability to use NO
3 as alternative
electron acceptor among a collection of fluorescent pseudomonads showed
that strains were either denitrifiers, NO
3 reducers, or not capable to respire
NO
3 (Clays-Josserand et al., 1995). Among the phygenetically diverse
group of denitrifiers, it is interesting that several bacteria are also involved
in other steps of the nitrogen cycle, such as nitrification or nitrogen fixation.
Thus, ammonia-oxidizing strains belonging to either the Nitrosospira or
Nitrosomonas genus have been shown to be capable to denitrify (Shaw
et al., 2006). It is also worth to note that the newly discovered group of
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Laurent Philippot et al.
Table 1 List of archaeal, bacterial, and fungal genera for which at least one
denitrifying strain has been characterized
Genus
Archaea
Haloarcula
Halobacterium
Pyrobaculum
Bacteria
Firmicutes
Bacillus
Paenibacillus
Actinomycetes
Corynebacterium
Streptomyces
Bacteroides
Flavobacterium
Example of species
Source
marismortui
denitrificans
aerophilum
(Yoshimatsu et al., 2000)
(Tomlinson et al., 1986)
(Völkl et al., 1993)
azotoformans,
stearotermophilus
terrae
(Ho et al., 1993; Pichinoty et al.,
1976b)
(Horn et al., 2005)
nephridii
thioluteus, sp.
(Har et al., 1965)
(Chèneby et al., 2000; Shoun et al.,
1998)
sp., denitrificans
(Horn et al., 2005; Pichinoty et al.,
1976a)
( Jones et al., 1992)
Flexibacter
canadiensis
Aquifaceae
Hydrogenobacter
thermophilus
Proteobacteria Alphaproteobacteria
Agrobacterium
sp.
Azospirillum
lipoferum
Bradyrhizobium
sp., japonicum
Brucella
melitensis
Hyphomicrobium
sp.
Mesorhizobium
loti
Ochrobactrum
anthropi
Paracoccus
pantotrophus
Pseudovibrio
denitrificans
Rhizobium
sp.
Rhodobacter
sphaeroides
Rhodopseudomonas salustris
Sinorhizobium
meliloti
Betaproteobacteria
Acidovorax
sp.
Alcaligenes
Achromobacter
Aquaspirillum
Azoarcus
faecalis
sp.
magnetotacticum
tolulyticus, anaerobius
(Suzuki et al., 2006)
(Chèneby et al., 2000)
(Neyra et al., 1977)
(Monza et al., 2006; van Berkum
and Keyser, 1985)
(Baek et al., 2004)
(Sperl and Hoare, 1971)
(Monza et al., 2006)
(Kim et al., 2006)
(Robertson and Kuenen, 1983)
(Shieh et al., 2004)
(Arrese-Igor et al., 1992)
(Sabaty et al., 1994)
(Kim et al., 1999)
(Daniel et al., 1982)
(Heylen et al., 2006; Schloe et al.,
2000)
(Vanniel et al., 1992)
(Youatt, 1957)
(Bazylinski and Blakemore, 1983)
(Fries et al., 1994; Springer et al.,
1998)
(continued)
257
Soil Denitrifiers
Table 1
(continued)
Genus
Example of species
Source
Azonexus
Azospira
Azovibrio
Burkholderia
Chromobacterium
Comamonas
caeni
sp.
sp.
sp.
sp.
sp., denitrificans
Cupriavidus
Dechloromonas
Denitratisoma
Kingella
necator
denitrificans
oestradiolicum
denitrificans, sp.
Microvirgula
Neisseria
Nitrosomonas
aerodenitrificans
sp.
europaea, eutropha
(Quan et al., 2006)
(Heylen et al., 2006)
(Heylen et al., 2006)
(Chèneby et al., 2000)
(Grant and Payne, 1981)
(Gumaelius et al., 2001; Patureau
et al., 1994)
(Pfitzner and Schegel, 1973)
(Horn et al., 2005)
(Fahrbach et al., 2006)
(Grant and Payne, 1981; Snell and
Lepage, 1976)
(Patureau et al., 1998)
(Grant and Payne, 1981)
(Poth and Focht, 1985; Zart and
Bock, 1998)
(Springs et al., 2004)
(Stamper et al., 2002)
(Magnusson et al., 1998)
(Tarlera and Denner, 2003)
(Schloten et al., 1999; Song et al.,
1998)
(Hole et al., 1996)
Ottowia
Ralstonia
Rubrivivax
Sterolibacterium
Thauera
thiooxydans
basilensis
sp.
denitrificans
aromatica,
mechernichensis
Thibacillus
denitrificans
Gammaproteobacteria
Halomonas
desiderata, campisalis
Luteimonas
Pseudomonas
mephitis
fluorescens, sp.
Pseudoxanthomonas taiwanensis
Shewanella
putrefaciens,
denitrificans
Stenotrophomonas
nitritireducens
Thioalkalivibrio
denitrificans
Zobellella
denitrificans,
taiwanensis
Epsilonproteobacteria
Nitratifractor
salsuginis
Nitratiruptor
tergarcus
Thiomicrospira
denitrificans
Eukaryota
Fungi
Fusarium
oxysporum
(Berendes et al., 1996; Mormile
et al., 1999)
(Finkmann et al., 2000)
(Gamble et al., 1977; Philippot
et al., 2001)
(Chen et al., 2002)
(Brettar and Hofle, 1993)
(Finkmann et al., 2000)
(Sorokin et al., 2001)
(Lin and Shieh, 2006)
(Nakagawa et al., 2005)
(Nakagawa et al., 2005)
(Brettar et al., 2006)
(Tanimoto et al., 1992)
258
Laurent Philippot et al.
ammonia oxidizers within the chrenoarcheota, possess the nirK gene encoding the denitrification NO
2 reductase (Treush et al., 2005), which suggests
that they can perform at least one step of the denitrification pathway.
Similarly, many nitrogen-fixing rhizobia can denitrify (Daniel et al., 1980,
1982; O’Hara and Daniel, 1985; van Berkum and Keyser, 1985). Even
though the diversity of denitrifiers is very high, it is likely that several yet
unknown microorganisms in nature contribute to the overall denitrification. As an example, Risgaard-Petersen et al. (2006) demonstrated that a
benthic foraminifer Globobulimina pseudospinescens accumulates intracellular
NO
3 stores, which can be respired to dinitrogen gas.
4. Assessing Denitrifiers Density, Diversity,
and Activity
4.1. Measuring denitrification and N2O emissions
Since denitrification is responsible for the loss of available NO
3 for plants,
many methods have been developed to estimate denitrification rates in soils.
The most basic approach calculates denitrification losses from the nitrogen
balance budget. However, other processes such as leaching can lead to NO
3
losses, which result in an overestimation of denitrification. An alternative
approach is based on the determination of the amount of N2O and/or
dinitrogen gas emitted by denitrification using various methods described
in the following sections.
4.1.1. Acetylene inhibition method
In this approach, acetylene (C2H2) is used to inhibit N2O reduction so that
total denitrification losses (N2 þ N2O) can be measured as N2O. The
blockage of N2O reduction in soil is obtained in an atmosphere containing
0.1–10% (v/v) C2H2. This method developed independently by Balderston
et al. (1976) and Yoshinari et al. (1977) has been a revolutionary key step in
estimating denitrification rates and has paved the way for hundreds of studies
measuring denitrification rates in situ (Stevens and Laughlin, 1998; Tiedje
et al., 1989). The C2H2 inhibition method has been applied to soil slurries
and cores (Ryden et al., 1987), as well as in field measurements using closed
chambers (Ryden and Dawson, 1982). For the latter, chambers are placed on
the soil surface and C2H2 is injected, which results in the accumulation of
N2O in the headspace of the chamber. The production of N2O is estimated
by analyzing gas samples from the headspace with a gas chromatograph,
preferably equipped with an electroncapture detector. The method has some
limitations related to the diffusion of C2H2 in soil, C2H2 degradation by
bacteria, and inhibition of other processes, for example, nitrification
(Keeney, 1986; Rolston, 1986).
Soil Denitrifiers
259
A widely used ex situ assay based on C2H2 inhibition has been developed
to measure the N2O production rate from the pool of active or activateddenitrification enzymes in a sample at the time of sample collection (Smith
and Tiedje, 1979b; Tiedje, 1982). This assay, termed the denitrifying
enzyme assay, is performed by incubating soil slurries under nonlimiting
denitrifying condition (i.e., no oxygen, saturating NO
3 concentration, and
addition of a surplus of electron donors). To avoid de novo enzyme synthesis,
samples are either incubated during a short period of time or in presence of
chloramphenicol, which blocks protein synthesis. The rate of N2O production, which is positively correlated to the amount of denitrification enzymes
in the samples, is then measured. As an alternative, the assay can be used
without addition of chloramphenicol and the denitrification rate can be
estimated by nonlinear regression (Pell et al., 1996). These assays can be used
to compare the effect of agronomical treatments on denitrification.
However, it does not provide information on field rates.
4.1.2. The isotope N-labeled methods
Denitrification activity can be determined using stable nitrogen isotopes in
both laboratory incubations and in field measurements. With this approach,
one or several 15N-labeled nitrogen compounds, such as NO
3 , ammonium,
fertilizers, or plant litter, are added to the soil. The subsequent production
dinitrogen and N2O by denitrification is measured by quantifying the
increase of 15N-labeled gases by mass spectrometry. As with the C2H2
inhibition method, closed chambers are used to estimate denitrification
activity in the field (Nason and Myrold, 1991). This method is limited by
the high cost of 15N and the need to add nitrogen in the soil. Methods based
on the use of 13N have also been described (Smith et al., 1978; Tiedje et al.,
1979), but these cannot be applied in the field (Tiedje et al., 1989).
4.2. Resolving diversity of denitrifiers
Over several decades, diversity of denitrifiers in soil was studied by isolating
bacterial strains. Basically, dilutions of soil suspension were spread on various
agar medium supplemented with NO
3 . After incubation under anaerobic
conditions, isolated colonies were characterized using phenotypic or metabolic tests, and later on by using molecular approaches (Chèneby et al., 2000,
2004; Garcia, 1977; Pichinoty et al., 1976a,b). The most complete survey
was reported by Gamble et al. (1977). From 19 soils, 3 freshwater lake
sediments, and oxidized poultry manure, around 1500 bacteria were isolated
and characterized. The dominant denitrifier populations in most samples
were related to Pseudomonas fluorescens. However, these isolation-based
techniques are limited by the fact that only a fraction of the bacterial
community is cultivable. Research on microbial diversity was completely
revolutionized 20 years ago by the application of molecular methods to
260
Laurent Philippot et al.
explore microorganisms in the environment without including a cultivation
step. These culture-independent molecular approaches have then been used
to assess the composition of denitrifier communities in soils.
The most frequently used approaches today to target denitrifiers in soil
start with extraction of nucleic acids (DNA or RNA) from the soil (Fig. 3).
The extracted nucleic acids are then purified and amplified by PCR using
primers targeting the denitrifier community. Since the ability to denitrify is
sporadically distributed both within and between different genera, and
cannot be associated with any specific taxonomic group, a 16S rRNAbased approach is not possible to target denitrifiers. However, in the late
1990s, the genes nirS and nirK encoding the key enzymes of the denitrification pathway were first used as molecular markers to describe the diversity
of the denitrifier community (Braker et al., 1998; Hallin and Lindgren,
1999). Since then, this approach has been extended to all the denitrification
genes (Braker and Tiedje, 2003; Flanagan et al., 1999; Philippot et al., 2002;
Scala and Kerkhof, 1999). Amplification of extracted nucleic acids using
primers targeting the denitrification genes is actually the most common way
to analyze denitrifier communities (Bothe et al., 2000; Hallin et al., 2007;
Philippot and Hallin, 2005, 2006). The sequence polymorphism of the
obtained mixed pool of PCR amplicons should reflect the composition of
the denitrifier community in the studied environment. The mixture of
PCR amplicons is analyzed by separating them based on their nucleotide
sequence polymorphism using either clone libraries combined with
sequencing or by fingerprinting techniques (Bothe et al., 2000; Hallin
et al., 2007; Philippot and Hallin, 2006). The most commonly used fingerprinting techniques to study denitrifier communities are terminal restriction
fragment length polymorphism (T-RFLP), restriction fragment length
polymorphism (RFLP), and denaturing gradient gel electrophoresis
(DGGE). These cultivation-independent approaches have limitations
related to the nucleic acids extraction, the choice of PCR primers, and
the PCR itself (Martin-Laurent et al., 2001; Philippot and Hallin, 2005).
4.3. Quantification of denitrifiers
Denitrifiers were first quantified by plating serial dilutions of soil suspension
and counting true denitrifying isolates based on their ability to reduce NO
3
into gaseous nitrogen production. However, the most common way to
count denitrifiers using a cultivation technique is to apply the most probable
number (MPN) method (Volz, 1977). Serial dilutions of soil suspension are
inoculated into anaerobic replicates medium tubes amended with NO
3 and
C2H2. Dilution tubes are then scored positive when N2O is detected, and
results are then converted into cell numbers copy using the McCrady table.
These methods refer only to microorganisms that can be cultivated and
therefore underestimate the actual number of denitrifiers in the sample.
261
Soil Denitrifiers
To circumvent this problem, molecular methods have also been developed
to quantify this functional community (Cho and Tiedje, 2002; Gruntzig
et al., 2001; Mergel et al., 2001; Michotey et al., 2000; TaroncherOldenburg et al., 2003; Tiquia et al., 2004; Ward et al., 1993). Two reviews
of these quantitative methods have been published (Philippot, 2006; Sharma
et al., 2007). Today, quantitative PCR is the main method used in soil
environments (Henry et al., 2004, 2006; Kandeler et al., 2006; LopezGutierrez et al., 2004; Qiu et al., 2004) (Fig. 3) with the same bias as for
the cultivation-independent approach for resolving community structure
outlined earlier.
1.000 E+1
1.000
Density analysis
1.000 E-1
1.000 E-2
1.000 E-3
1.000 E-4
1.000 E-5
0
5
10
15
20
25
30
35
40
Real time-PCR
Competitive-PCR
Quantitative-PCR
Nucleic acids
extraction
Structure analysis
Soil samples
Fingerprint analysis
Amplification
Clone-library analysis
50 100 150 200 250 300 350 400 450 500 550 600 650
T-RFLP
RFLP
DGGE
Sequencing
RFLP
Figure 3 Methods used to assess diversity and density of denitrifiers with a PCR-based
approach.
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Laurent Philippot et al.
5. Natural Factors Causing Variations in
Denitrification
5.1. Temperature and water
Both the overall denitrification rates and the proportions of N2O and
dinitrogen gas produced by denitrifying microbes can vary depending on
numerous environmental factors, such as pH, carbon, NO
3 , and NO2
availability, soil moisture, pore structure, aeration, temperature, freezing–
thawing, and drying–wetting events. Several of these are natural factors
influenced by climatic conditions that cannot be managed. In addition,
they are not constant, but show large variation over the vegetation period
as well as between field sites. The estimated nitrogen losses are therefore
highly variable in time and space. Emissions of N2O and dinitrogen show no
consistent seasonal pattern. In some studies, the largest N2O emissions were
recorded during spring (Kaiser and Heinemeyer, 1996; Parsons et al., 1991;
Ryden, 1985), in others during spring and autumn (Ambus and Christensen,
1995; De Klein and Van Logtestijn, 1994), or in summer (Bremner et al.,
1980; Cates and Keeney, 1987). The difference in the results could not be
related to environmental factors and management practices. A better understanding of factors contributing to variability of denitrification activity
would be helpful to improve estimations and modeling of nitrogen fluxes
by denitrification.
Soil temperature and soil water content are known factors that affect
gaseous nitrogen losses and the N2O/N2 ratio. Under constant laboratory
conditions, this ratio increased exponentially with increasing soil temperature (Maag and Vinther, 1996). However, the ratio was strongly influenced
by soil type, although these data could not be confirmed by field measurements. Whereas Bailey (1976) and McKeeney et al. (1979) found a positive
correlation between soil temperature and denitrification activity, others
observed no relationship with temperature (Focht, 1974; Lensi and
Chalamet, 1979). The reason might be the lower water content caused by
increased plant transpiration rates at higher temperatures, which leads to a
water deficiency. Under laboratory conditions, similar to the effects of
increasing temperature, the overall denitrifying activity and N2/N2O ratio
increased with increasing soil water content (Colbourne and Dowdell,
1984; Vinter, 1984). This was also confirmed in a pasture after harvest
(Rudaz et al., 1997).
Linked to soil water content is oxygen availability. Hochstein et al.
(1984) showed that soil oxygen concentrations below 5% resulted in denitrification being the main microbial respiratory process when NO
3 was
available. In addition, at 10% oxygen concentration and moisture content
between 40% and 60%, denitrification was the main source of emitted N2O.
Soil Denitrifiers
263
Water content depends on the pore structure of the soil, which in turn is
affected by soil type, organic matter content, and land use. Bakken et al.
(1987) demonstrated that the pore space structure appears to be the major
factor explaining the difference in mean denitrification rates by comparing
pasture and cropped soil. In the field, Bijay-Singh et al. (1989) found higher
actual denitrification in cropped soil than in pasture, despite similar NO
3
contents. They explained their results as the consequence of better drainage
in the pasture soil, due to the higher porosity of this soil. Complementary
measurements after the application of various amounts of water showed
denitrification activity in pasture soil was higher than denitrification in
cropped soil only at water suctions greater than 5.5 kPa (Bijay-Singh
et al., 1989). In contrast, potential denitrification has often been reported
to be higher in pasture than in cropped soil (Bijay-Singh et al., 1989; Lensi
et al., 1995; Sotomayor and Rice, 1996).
5.2. Freeze–thaw cycles
5.2.1. Freeze–thaw effects on nitrous oxide emissions
Christensen and Tiedje (1990) were the first to report peak N2O emissions
from arable soils in spring during thaw periods. Emissions of carbon dioxide
and N2O and uptake of methane throughout the snow-covered period even
at temperatures near 0 C were later reported (Sommerfeld et al., 1993).
In order to decide whether N2O production can be attributed also to
nonmicrobial processes in soil, emissions from a g-ray sterilized and a
nonsterilized soil were compared in a laboratory experiment, where the
freezing and thawing cycles were simulated. The results clearly indicated
that microbial processes were responsible for N2O production in thawing
and even frozen soils (Röver et al., 1998). Therefore, efforts have been done
to investigate the effects of freezing and thawing cycles on microbial
denitrification, and to understand the mechanisms behind. Sehy et al.
(2003) first demonstrated the importance of denitrification for nitrogen
losses during winter in arable soil. They separated the 12 months of investigation into the growing season (March to November) and the winter period
(December to February). Independent of the amount of applied fertilizer,
about 70% of the annual N2O amounts was emitted during the winter
period. The temporal changes of the N2O emission rates were correlated to
changes in soil temperature. Similarly, Dörsch et al. (2004) found persistently high N2O emissions in arable soil with peak emissions during midwinter thawing, diurnal freezing–thawing, and spring thaw. Low and stable
temperatures below the insulating snow or ice cover, in contrast, decreased
N2O emissions. Several other field studies in the temperate regions also
reported high N2O emissions from agricultural soils during freeze–thaw
periods reaching 20–70% of the annual budget (Flessa et al., 1995; Nyborg
et al., 1997; van Bochove et al., 1996, 2000; Wagner-Riddle et al., 1997).
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Laurent Philippot et al.
Nevertheless, a few studies have also reported that moderate freeze–thaw
fluctuations had little impact on nitrogen dynamics and N2O emissions in
soils (Grogan et al., 2004; Neilsen et al., 2001).
There is considerable debate on which factors could be critical controllers of winter N2O emissions from arable soils. However, most authors state
that emissions during winter are related to the release of nutrients.
Christensen and Christensen (1991) could show that soluble carbon, applied
as plant extract, was necessary to induce N2O production during freezing
and thawing events. Therefore, plant residues from catch crops and green
manure may play an important role in the regulation of N2O emissions in
winter, since frost enhances the release of organic compounds from plant
residues. Additionally, freeze–thaw events may result in transient pulses of
carbon and nitrogen due to disruption of soil aggregates (Christensen and
Christensen, 1991; Müller et al., 2002) and lysis of microorganisms (Schimel
and Clein, 1996; Skogland et al., 1988). Müller et al. (2002) showed that the
increased ammonium and NO
3 concentrations during freezing were associated to peak N2O emissions during the following thawing period.
Enhanced oxygen consumption during degradation of plant residues combined with a high water content of the thawing soil increases the anaerobic
volume, thus enhancing denitrification. The freeze–thaw-induced emission
of N2O could thus be a straightforward result of enhanced denitrification.
N2O may also be produced by microorganisms in unfrozen water films on
the soil matrix during freezing. Several authors showed that an ice layer
covering the unfrozen water film could be a diffusion barrier, which reduces
oxygen supply to the microorganisms and partly prevents the release of
N2O to the air (Burton and Beauchamp, 1994; Goodroad and Keeney,
1984; Teepe et al., 2001).
Nitrification could also be of significance for N2O emissions during
winter. It has been demonstrated that freeze–thaw cycles enhances nitrogen
mineralization, which results in the release of substrate for ammoniaoxidizing bacteria (Deluca et al., 1992). Lowered oxygen availability during
freeze–thaw-induced respiration could also induce higher N2O emissions
from nitrifiers, since the N2O/(NO
3 þ NO2 ) ratio of nitrification increases
sharply in response to oxygen limitation (Davidson, 1991; Dundee and
Hopkins, 2001; Goreau, 1980). However, it has been demonstrated that
only a few percent of the measured N2O originate from nitrification.
Denitrification was the main N2O source at various oxygen concentrations
investigated in freeze–thaw-affected soil (Ludwig et al., 2004; Mrkved et al.,
2006).
5.2.2. Freeze–thaw effects on denitrifier communities
Although microbial denitrification is believed to be the major source of
N2O during freeze–thaw events, few have analyzed the denitrifier communities involved. Actually, little is known about the significance of the
Soil Denitrifiers
265
denitrifier community composition for N2O emissions in general, since
most of the work conducted has focused on gas and soil analysis. Freeze–
thawing effects on total bacterial community structure are contradictory.
Eriksson et al. (2001) observed a change in ribosomal internal spacer analysis
patterns during freeze–thaw events, whereas Koponen et al. (2006) concluded that neither microbial biomass nor community structure was affected
in boreal soils.
It has been postulated that the relative activity of N2O reductase can be
lowered at near-freezing temperatures (Holtan-Hartwig et al., 2002b; Melin
and Nömmik, 1983), possibly resulting in high N2O/(N2 þ N2O) ratios in
soil during thawing. A high N2O/(N2 þ N2O) ratio could also be a
‘‘postfreezing trauma’’; the N2O reductase appears to be more vulnerable
to perturbations than the other denitrification enzymes, and if this holds
for frost damages, it would result in a higher proportion of produced N2O
to total denitrification after freezing (Dörsch and Bakken, 2004; HoltanHartwig et al., 2002; Melin and Nömmik, 1983). Nevertheless, how specific
enzymes involved in denitrification are influenced by freezing and thawing
is still not answered.
Sharma et al. (2006) investigated the mRNA levels of genes encoding
the periplasmic NO
3 reductase gene (napA) and cytochrome cd1 NO2
reductase (nirS) in the upper horizon of a grassland soil during thawing in
a laboratory experiment. By using a MPN-based reverse transcriptase PCR
approach they could show that high transcript levels occurred for both
genes 2 days after thawing had begun, followed by a decrease. The peak
of N2O production coincided with the peak for napA and nirS transcripts,
and it timely shifted after 2 days. In the same study, the napA and nirS
genotype diversity was analyzed. Interestingly, DNA-based profiles showed
no change in banding patterns, whereas those derived from cDNA showed
a clear succession of the genotypes, with the most diverse community
structure at the time point of the highest gene expression.
5.3. Dry–wet cycles
Similar to freeze–thaw cycles in soil, dry–wet cycles can enhance N2O
emissions. Priemé and Christensen (2001) compared the effects of drying–
wetting and freezing–thawing cycles on the emission of N2O, carbon
dioxide, and methane from intact soil cores from farmed organic soils.
During the first week, following wetting or thawing, up to a 1000-fold
increase in N2O emission rates were recorded from the cores. The total N2O
emission ranged between 3 and 140 mg N–N2O m2, and between 13 and
340 mg N–N2O m2 due to the first wetting and thawing event, respectively. Nevertheless, the emission rates declined after two successive freeze–
thaw events. Many other studies have also documented differences in the
rate of denitrification following wetting (Ambus and Lowrance, 1991;
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Laurent Philippot et al.
Gilliam et al., 1978; Groffman and Tiedje, 1989; Rice and Tiedje, 1982;
Robertson and Tiedje, 1985, 1988; Sexstone et al., 1986). Some studies have
also noted denitrification differences between the wet-up and dry-down
phases of soil moisture following rainfall events (Gilliam et al., 1978).
Bergsma et al. (2002) showed that a short wet-treatment significantly
decreased the relative amount of N2O emitted from cropped soil compared
with a long wet-treatment, while no effect of moisture history was seen in a
successional agrosystem. The authors hypothesized that these differences in
N2O production were due to selection of denitrifiers with enhanced capacity for enzyme maintenance at lower levels of NO
3 , such as found in the
successional soil. Others later confirmed differences in denitrifier community composition in the successional and cropped soil at this site (Stres et al.,
2004). Denitrification enzymes were also more sensitive to oxygen in the
cropped soil and N2O activity was higher in the successional soil (Cavigelli
and Robertson, 2000). Soil moisture history seems to be important for
denitrification. If denitrification enzymes are induced differentially in
response to wetting, then both the overall rate of denitrification as well as
the relative amount of N2O will differ substantially among ecosystems.
6. Denitrification in the Rhizosphere of Crops
6.1. Crops as a factor influencing denitrifiers
The rhizosphere is the volume of soil influenced by plant roots (Hiltner, 1904).
The growth and activity of the root system induce significant modifications in the physicochemical and biological properties of the soil surrounding
the roots, which correspond to the so-called rhizosphere effect. It is well
known that the major factors regulating denitrification: carbon, oxygen, and
NO
3 can be modified in the rhizosphere of plants. Thus, carbon compounds,
which can be used as electron donor by denitrifiers, are released by plants roots
in the surrounding soil through rhizodeposition. The effect of plants on
oxygen and NO
3 concentration is more complex. Oxygen concentration
can be lowered in the rhizosphere by respiration of the roots and microorganisms. On the other hand, consumption of water by plant roots increases soil
gas exchange and oxygen concentration. Some plants, such as rice, also
transport oxygen from the air down to the soil in water-saturated soil. Finally,
when roots grow and penetrate the soil, they can modify soil compaction,
which affects oxygen diffusion. Nitrate is used by both plants and microorganisms and the competition for NO
3 is therefore high in the rhizosphere during
the growing season. However, plants can also potentially provide NO
3 for
denitrification when organic matter present in root exudates is mineralized.
Moreover, during plant senescence and litter decomposition in fall and
winter, nitrogen becomes bioavailable and can be denitrified. Overall, factors
Soil Denitrifiers
267
regulating denitrification in the rhizosphere are strongly interwoven and the
stimulating effect of root-derived carbon is only observed under nonlimiting
concentrations of NO
3 and oxygen. It is therefore not possible to state that
plant roots always stimulate denitrification.
6.1.1. Effect of crops on the denitrification activity
Comparison of denitrification rates between planted and nonplanted soil in
the field or in incubation experiment has been the most common approach
to investigate the influence of crops on this process. Early reports showed
enhanced denitrification rates in the rhizosphere compared with bulk soil
(Smith and Tiedje, 1979a; Stefanson, 1972; Woldendorp, 1962). The key
role of plant on denitrification has later been confirmed in several studies,
although the mechanisms responsible for the higher denitrification rates are
still not clear. Among the agricultural plants studied, barley (Hordum vulgare)
has received the greatest attention so far. Klemedtsson et al. (1987) observed
that denitrification rates in pots planted with barley increased with time
along with increased root biomass. Stimulation of the denitrification rates in
planted pots was 2–22 times compared with the unplanted pots. Similar
results were reported by Hjberg et al. (1996) who observed an average
NO
3 reduction and denitrification rates in the rhizosphere of barley
1.8 times higher than in the bulk soil, with the most pronounced increase
of 7 times. By using monoclonal antibodies against the copper nitrite
reductase, Metz et al. (2003) clearly showed the presence of active enzymes
in the rhizosphere of wheat.
Vinter et al. (1984) demonstrated that this increase of denitrification in
the barley rhizosphere was positively correlated with soil NO
3 concentration.
Their results showed that for fertilizer applied to barley at 30 kg N ha–1, the
denitrification rate increased 2.5 times while a fivefold increase was observed
in field plots receiving 150 kg N ha–1. These results were consistent with those
of Mahmood et al. (1997), who carried out a field experiment to examine the
–1
effect of maize plants on denitrification. At low soil NO
3 levels (1–4 mg N g
dry soil), the presence of maize plants resulted in a nearly 50% increase in
–1 dry soil) the
denitrification, whereas at higher NO
3 levels (7–19 mg N g
observed increase due to plants was 2.5 times. The combined effect of
plant roots and NO
3 concentration on denitrification was first pointed out
by Smith and Tiedje (1979a). They found that denitrification was lower in
planted than in unplanted soil when NO
3 concentration was low (0.002 g
1 dry soil), while at higher NO concentration (0.1 g NO –N
NO
–N
kg
3
3
3
kg1 dry soil) the presence of plants increased denitrification. Qian et al. (1997)
also reported higher denitrification rates in the unplanted soil compared
with planted soil at late maize growth stages when the amount of NO
3 was
limiting in the planted soil. These neutral or negative effects of plant roots on
denitrification were attributed to NO
3 depletion around the roots.
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Laurent Philippot et al.
It has also been reported that the rhizosphere effect on denitrification
was associated with air-filled porosity (Wollersheim et al., 1987). At a low
moisture tension, Bakken (1988) observed a tenfold increase in the denitrification rate in the planted soil compared with the unplanted soil. At
medium or high moisture tension, the plants had no or even a negative
effect on denitrification. Similarly, Prade and Trolldenier (1988) reported
that the rhizosphere effect on denitrification was confined to air-filled
porosity lower than 10–12% (v/v). Thus, the lack of stimulation on denitrification in the rhizosphere at nonlimiting NO
3 concentrations reported
by Haider et al. (1985) was attributed to a high air-filled porosity in both
planted and unplanted pots.
Carbon, the third factor regulating denitrification, is probably responsible for the stimulating effect of plants on denitrification activity. Several
investigators have demonstrated the influence of different organic substrates
on denitrification. Denitrification was correlated with soluble organic matter (Bijay-Singh et al., 1988; Burford and Bremner, 1975; Cantazaro and
Beauchamp, 1985; McCarty and Bremner, 1993) and easily mineralizable
carbon (Bijay-Singh et al., 1988). The release of organic compounds by
living roots can directly affect denitrification rates by providing an additional source of electron donor, but also indirectly by increasing microbial
activity, which lowers the oxygen concentration. This amount of carbon
released by roots into the soil can be up to 20% of photosynthetically
fixed carbon during the vegetation period (Hütsch et al., 2002; Nguyen,
2003). The nature of the root-derived carbon is highly variable (mucilage,
exudates, root cap cells, and so on). The mucilage is composed of highmolecular-weight polysaccharides, mainly arabinose, galactose, fucose,
glucose, and xylose, and up to 6% is proteins. In contrast, exudates are
low-molecular-weight compounds released passively from roots such as
sugars, amino acids, and organic acids. As expected, daily addition of
70 mg C g–1 dry soil of maize mucilage to an agricultural soil increased
denitrification 2.8 times compared with water addition (Mounier et al.,
2004). Similarly, daily addition at a rate of 150 mg C g–1 dry soil of different
mixtures of amino acids, organic acids, and sugars mimicking maize root
exudates greatly stimulated denitrification rates (Henry et al., unpublished
data). In addition, several investigations have shown that denitrification
rates were also positively related to the distribution of fresh plant residues
in the soil profile (Aulakh et al., 1984, 1991; Cantazaro and Beauchamp,
1985; Christensen and Christensen, 1991; Parkin, 1987).
6.1.2. Effect of crop on the denitrifier community
In contrast to denitrification activity, there have been fewer studies of the
effect of plant on the denitrifier community. Vinther et al. (1982) reported
some early estimates of the diversity and the density of denitrifiers in
agricultural soils under continuous barley cultivation. Counts of denitrifiers
Soil Denitrifiers
269
performed using the most-probable-number method with NO
3 agar broth
as growth medium revealed densities ranging between 103 and 106 bacteria
g1 of dry soil, which represented less than 1% of total bacteria. In contrast,
NO
3 reducers counts for less than 10% of total viable count. Identification
of denitrifying isolates based on selected physiological and morphological
properties showed that numerically predominant denitrifiers belonged to
Pseudomonas spp., Alcaligenes sp., and Bacillus sp. The effect of plant roots on
the taxonomic diversity of denitrifiers has further been investigated by
isolating denitrifiers from unplanted or maize planted soil in a 3-month
incubation experiment (Chèneby et al., 2004). Density of denitrifiers was
2.4 106 and 1.6 107 cells g1 of dry soil in the unplanted and planted
soil, respectively. A total of 3240 NO
3 -reducing isolates were obtained and
188 of these isolates were identified as denitrifiers based on their ability to
reduce at least 70% of the NO
3 to N2O or N2. Comparison of the
distribution of the denitrifying isolates between planted and unplanted soil
showed a difference in the composition of the denitrifier community with
an enrichment of phylogenetically Agrobacterium-related denitrifiers in the
planted soil. In addition, these predominant Agrobacterium-related isolates
from the rhizosphere soil were not able to reduce N2O while dominant
isolates from the unplanted soil emit N2 as end denitrification product.
Direct molecular approaches have recently been applied to investigate
the effect of maize on NO
3 reducers community performing the first step of
the denitrification pathway. The narG gene encoding the membrane-bound
NO
3 reductase was used as molecular marker to analyze the composition of
the NO
3 reducers community from planted and unplanted pots after
3 months of repeated maize culture. A shift in the community composition
between unplanted and planted soils was reported without significant modification of the diversity indices (Philippot et al., 2002b). Clone library
analysis revealed that most of the dominant sequences in the planted soil
were related to narG from the Actinomycetes suggesting a specific selection
of NO
3 -reducing actinobacteria by the maize roots. In contrast, Chèneby
et al. (2003) detected a reduction of the reciprocal Simpson’s diversity index
in the maize planted soil compared with the unplanted soil, but without any
major modification of the composition of the NO
3 -reducing community
in another soil type. The results from these two studies suggest that the
rhizosphere effect on the structure of the denitrifier community is strongly
dependent on the soil type. Several studies aiming at sorting out the relative
importance of plant and soil confirmed that these two factors might act
simultaneously in determining the composition of the indigenous soil
microbial community (Clays-Josserand et al., 1999; Costa et al., 2006;
Marschner et al., 2004; Wieland et al., 2001).
In two studies, effort has been devoted to disentangle the mechanism of
the rhizosphere effect by investigating the influence of the two major
rhizodeposits, mucilage and exudates, on the genetic structure of denitrifiers
270
Laurent Philippot et al.
(Henry et al., unpublished data; Mounier et al., 2004). Analysis of the
structure of the denitrifier community by direct molecular approaches
revealed only minor changes after mucilage amendment (Mounier et al.,
2004). Similarly, the addition of sugar, amino acids, and organic acids
mimicking maize exudates resulted in minor changes in the structure and
the density of the denitrifier community (Henry et al., unpublished data).
Even though root-derived carbon can stimulate denitrification activity, it
does not seem to be an important driver of the denitrifier community
structure in soil. However, the community structure of the active members
of the denitrifying community might be influenced by root exudates, but
this has not yet been clarified.
6.1.3. Denitrification provides a selective advantage in
the rhizosphere
Since most of denitrifiers are chemoheterotrophs, the increase of denitrifier
density together with total microbial density observed in the rhizosphere
was mainly attributed to the higher availability of organic substrates in the
root vicinity. However, it has been suggested that the ability to grow by
respiring nitrogenous compounds when oxygen is limited could be a
selective advantage for denitrifiers in the rhizosphere. Thus, using DNA
probes for the gene encoding the NO
2 and N2O reductase, von Berg and
Bothe (1992) found that the denitrifier to other heterotrophic organism
ratio was increased near the roots. Such influence of plants on the distribution of denitrifying abilities has also been reported by Clays-Josserand et al.
(1995), who observed that the proportion of denitrifying pseudomonas
isolates gradually increased in the root vicinity of tomato. To demonstrate
that this selection of denitrifiers in the rhizosphere was due to ability
to respire nitrogenous and not to other traits, the competitive abilities of
denitrifying strains in the rhizosphere have been compared with those
of their isogenic nondenitrifying mutants. Mutants unable to synthesize
either the membrane-bound NO
3 reductase, the cd1 NO2 reductase, or
the copper nitrite reductase were outcompeted by the denitrifying wildtype strains in the rhizosphere of maize demonstrating that denitrification
itself could provide an advantage for root colonization (Ghiglione et al.,
2000; Philippot et al., 1995).
6.2. Impact of crop species, crop cultivars, and
transgenic plants
Because both shoot and root properties, for example, different litter types
and roots architecture, and the amount and composition of root exudates
are varying among plant species and cultivars (Hütsch et al., 2002), it has
been hypothesized that effect of plants on microorganisms differ depending
on plant species or cultivars. Therefore, in the last decade many studies were
Soil Denitrifiers
271
performed to prove this hypothesis. Most were based on 16S rRNA
approaches, which make it impossible to relate any changes in the microbial
community structure to functions. Only few attempts were made to use
functional genes to measure possible impacts of crop species or cultivars on
microorganisms involved in nitrogen cycling.
6.2.1. Rhizosphere effect on denitrification depends on crop species
Effects of crop species or cultivars have mainly been investigated on denitrification activity rather than on the diversity of denitrifiers. Crush (1998)
reported a tendency for higher potential denitrification rates in association
with bigger root mass in a lysimeters study with various forage plants.
Differences in the denitrification rates between small grains (barley,
wheat, and oats) and grasses were also reported by Bakken (1988). Since
legume plants associated with nitrogen-fixing bacteria can be used as substitute for mineral fertilizers, several authors studied whether their cultivation
affect the nitrogen cycle processes. Using the C2H2 inhibition technique on
intact soil cores sampled during 2 years in a field, Svensson et al. (1991)
reported significant differences between plant species with higher denitrification rates with lucerne (Medicago sativa L.) than with barley (Hordeum
disticum) and grass ley (Festuca pretensis Huds.). Larger denitrification rates
under legumes than other plants were also reported by other studies (Kilian
and Werner, 1996; Scaglia et al., 1985). The higher positive effect of legume
on denitrification rates was observed not only with living plants but also
during their decomposition process. Aulakh et al. (1991) and McKenney
et al. (1993) showed higher denitrification rates in soil amended with
legume residues than in soil amended with grass, corn, or wheat residues.
However, lower denitirification rates were observed with clover than with
small grains or grasses (Bakken, 1988).
It has been hypothesized that the higher denitrification rates caused by
legumes could be due to their symbioses with denitrifying Rhizobiacaea.
Thus, several studies reported that denitrification was very common in
rhizobia (Asakawa, 1993; Daniel et al., 1980, 1982; O’Hara and Daniel,
1985; Tiedje, 1988; van Berkum and Keyser, 1985; Zablotowicz et al.,
1978) and that many strains can denitrify both as nodule bacteroids and in
the free-living state (Arrese-Igor et al., 1992; Garcia-Plazaola et al., 1995).
Accordingly, Kilian and Werner (1996) showed that mean denitrification
was increased fourfold in plots of the nitrogen-fixing bean Vicia alba compared with nonnodulated V. alba mutant. On the other hand, GarciaPlazaola et al. (1993) suggested that even with optimal conditions for
denitrification and the highest rhizobial populations found in agricultural
soils, the contribution of Rhizobiacaea to the total denitrification was virtually neglectable as compared with other soil microorganisms. The fact that
different legume plants were analyzed may explain these contrasting results.
Since the symbiosis between rhizobia and legume plants is highly specific,
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Laurent Philippot et al.
different rhizobial strains, which can exhibit contrasted denitrification abilities, are selected according to the legume species. This hypothesis is
supported by the work of Sharma et al. (2005) who studied the diversity
of transcripts of the NO
2 reductase in the rhizosphere of three different
legumes: Vicia faba, Lupinus albus, and Pisum sativum. A significant plantdependent effect on the transcripts was observed, suggesting that the active
denitrifiers were different in the rhizosphere of three legumes. The denitrifier community structure, based on the DNA analysis of nirK and nirS genes,
was not as variable between the different plant rhizospheres, indicating a
stable denitrifier community. Similar results were also found by Deiglmayr
et al. (2004). When investigating the effect of Lolium perenne and Trifolium
repens on the NO
3 reducer community, based on DNA analysis of narG, no
plant species effect was observed. In contrast, with a similar approach, Patra
et al. (2006) observed an effect of the plant species on both the structure and
the activity of the denitrifier community among Arrhenatherum elatius,
Dactylis glomerata, and Holcus lanatus in grasslands.
6.2.2. Impact of transgenic crops
Transgenic crops offer agronomic advantages, such as improved yield,
improved product quality, herbicide tolerance, or insect resistance, over
their corresponding nontransgenic wild-type cultivar. These modifications
are mostly obtained by adding a gene in the genome of the parental wildtype crop via genetic manipulation. Plant genetic engineering can be beneficial when it improves agronomic features, but ethical concerns and the
impact of genetically modified crops on human health and on the environment is under debate. Therefore, quantitative risk assessments have been
undertaken to determine the safety of transgenic plants. Such studies were
performed on not only insects, earthworms, nematodes, and so on, but also
on microorganisms, which dominate soil-borne communities. Like plant
developmental stage or genotype can influence microbial diversity and
activity in the rhizosphere (Rengel et al., 1998), introduction of a transgene
might modify the plant effect on microorganisms, due to altered root
rhizodeposition (Kowalchuk et al., 2003). For example, Bacillus thuringiensis
toxins (Bt) produced by transgenic plants are released in the soil by root
exudates (Saxena et al., 1999), which possibly affects the soil microorganisms. Indirect effects of transgenic crops on soil microbes could arise from
repeated application of herbicide during cultivation of herbicide-resistant
plants (Sessitsch et al., 2004).
Most of the studies investigating effects of transgenic crops on soil
microorganisms have focused on total bacteria (Baumgarte and Tebbe,
2005; Heuer et al., 2002; Lukow et al., 2000; Milling et al., 2004;
Schmalenberger and Tebbe, 2002). However, Philippot et al. (2006) compared the effect of glyphosate-tolerant maize, treated with either glyphosate
or atrazine, and two cultivars of pyrale corn pest-resistant maize, treated
Soil Denitrifiers
273
with atrazine, on the NO
3 -reducing community in a field experimented
during 8 years. The nitrate reductase activity was higher in the rhizospheric
soil than in the bulk soil, but no difference between the three cultivars was
observed. A rhizosphere effect was also observed on the NO
3 -reducer
community structure together with a strong influence of the sampling
date, but the type of cultivar did not matter. Accordingly, analysis of the
NO
3 -reducing community structure in the rhizosphere of five different
cultivars of transgenic maize and the corresponding parental wild-type
cultivars in a greenhouse experiment did not reveal any transgene effect
(Sarr et al., unpublished data).
7. Impact of Fertilization on Denitrification
7.1. Fertilization affects denitrification
Research on denitrification in agricultural soil has mainly focused on effects
of fertilizers. Not surprisingly, nitrogen fertilizers promote denitrification
activity in agricultural soil and substantial amounts of fertilizer added nitrogen is lost through denitrification (De Klein and Van Logtestijn, 1994;
Kaiser et al., 1998; Mulvaney et al., 1997; Ryden, 1983). Fertilization can
also affect the N2O to N2 ratio from denitrification, and N2O emissions are
most likely increasing due to an increased input of fertilization (Skiba and
Smith, 2000). It has often been suggested that denitrification is limited
under field conditions by NO
3 availability (Bronson et al., 1992;
Mahmood et al., 2005), which in turn is influenced by the fertilizer type
and application rate together with timing and application method. For
example, losses by denitrification are often highest shortly after fertilization
application and these losses can account for 50–75% of the annual loss in a
field (Ellis et al., 1998; Mogge et al., 1999). The combination of high
nitrogen application rates and poor soil drainage give rise to higher denitrification activity than lower application rates and good drainage (Hofstra and
Bouwman, 2005). De Klein and Van Logtestijn (1994) showed that high
nitrogen losses were associated to soil water content rather than as an effect
of application rates in mineral fertilized grasslands. Fertilization sometimes
causes secondary effects that affect denitrification. Such secondary effects
can be changes in pH. Changes in pH can both directly and indirectly affect
denitrification activity, and in general, denitrification is higher at neutral
rather than acidic conditions (Bremner and Shaw, 1958; Nömmik, 1956;
Šimek and Cooper, 2002). Organic fertilizers can also cause secondary
effects on denitrification by the various organic and inorganic compounds
that are found in the fertilizers. For example, the high heavy metal content
occasionally found in sewage sludge can decrease denitrification.
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Laurent Philippot et al.
7.1.1. Inorganic and organic fertilizer effects on denitrification activity
The various ammonium-based fertilizers affect denitrification differently,
due to the effect the fertilizer has on soil pH. Some of these fertilizers
hydrolyze in soil, which gives an acidic reaction, while others are alkaline
forming. Not only denitrification, but also nitrification is higher at neutral or
alkaline compared with acidic conditions (Prosser and Embley, 2002) and,
therefore denitrification is additionally supported by the supply of NO
3
from the nitrifiers under these conditions. It is also known that alkaline
forming fertilizers affect the dissolution of organic matter (Norman et al.,
1987; Sen and Chalk, 1994), thus increasing the amount of solubilized
carbon and nitrogen that can be used for denitrification. Other microbial
processes also benefit from the released nutrients, which results in reduced
oxygen concentrations that promote denitrification. Accordingly, Mulvaney
et al. (1997) reported higher emissions of N2O and dinitrogen gas after
application of alkaline-hydrolyzing fertilizers than after application of
acidic fertilizers, with the following order: anhydrous NH3 > urea >>
(NH4)2HPO4 > (NH4)2SO4 NH4NO3 NH4H2PO4. In this laboratory
study, all the fertilizers tested promoted denitrification, but from a 20-yearold field experiment, Simek et al. (Šimek and Hopkins, 1999; Šimek and
Kalcik, 1998) reported that large amounts of a mix of different fertilizers
could decrease denitrification, in some cases even below the rates observed in
unfertilized soils, when no lime was applied. Results from a long-term field
trial showed that potential denitrification rates were much lower in plots
fertilized with ammonium sulfate, which had acidified the soil to pH 3.97,
compared with calcium nitrate fertilized plots having pH 6.26 (Enwall et al.,
2005) (Fig. 4). Similarly, application of potassium nitrate increased the rates
of denitrification more than an ammonium sulfate-based fertilizer in a
flooded subtropical soil (Aulakh et al., 2000).
Organic fertilizers often promote denitrification more than mineral
nitrogen fertilizers and this has been reported in numerous studies
(Dambreville et al., 2006; Ellis et al., 1998; Enwall et al., 2005; Magnusson
et al., 1998; Rochette et al., 2000; Wolsing and Priemé, 2004). Organic
fertilizers include the various types of farm manure commonly used, but also
green manures, crop residues, sewage sludge, composted wastes, and other
wastes. The stimulation of denitrification by organic fertilizers is probably
due to the additional supply of readily available organic carbon (Christensen,
1985). However, since organic fertilizers release nitrogen slowly, the supply
of nitrogen is initially low. This explains why some studies reported low
denitrification rates in organically fertilized soil compared with soils with
mineral fertilization the first years after application in new field experiments
(Estavillo et al., 1994, 1996; Schwarz et al., 1994).
Similarly to mineral fertilizer, the type of organic fertilizers influences
the denitrification rates. Different fertilizers by default contain different
Soil Denitrifiers
275
Figure 4 Long-term fertilization experimental field site established in 1956 at Ultuna
campus, Uppsala, Sweden.
nitrogen and carbon concentrations, as well as different amounts of inorganic and organic pollutants. They also differ in acidification capacity. All
these factors affect denitrification. For example, nitrogen losses by denitrification from a site fertilized with farmyard manure were twice those from a
site fertilized with cattle slurry, even though the nitrogen addition was three
times higher in the latter (Mogge et al., 1999). This could be explained by
the difference in C/N ratio, but an effect of pH and different crop rotation
history cannot be ruled out. It has also been reported that digested pig slurry
and composted pig slurry reduced the denitrification losses by 30% compared to untreated pig slurry (Vallejo et al., 2006). Others showed that
pretreatment affects both the nutrient status of the fertilizers and the amount
of and type of organic pollutants present, which affected nitrogen cycling in
soil (Levén et al., 2006; Nyberg et al., 2006). Long-term fertilization with
cattle manure was shown to increase potential denitrification rates compared with fertilization with sewage sludge, even though equal amounts
based on carbon content had been added and both the soil nitrogen and
carbon content was comparable between the treatments (Enwall et al.,
2005). It was argued that the lower pH itself caused by the sewage sludge
was not a sufficient explanation for the lower denitrification activity, and
elevated heavy metal concentrations were found in the sewage treated plots
(Bergkvist et al., 2003; Witter and Dahlin, 1995). In two other field
experiments, 12 and 16 years of sewage sludge application had positive
effects on soil potential denitrification, even though copper increased in the
soil and pH dropped slightly during this period ( Johansson et al., 1999).
Amendment with different crop residues has also been shown to affect
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Laurent Philippot et al.
nitrogen losses by denitrification differently (Velthof et al., 2002). Various
Brassicacaea species caused higher losses than residues from grasses, probably
due to the lower C/N ratios and higher amounts of mineralizable nitrogen
in the former crop residues.
Despite possible increased denitrification activity, declines in soil organic
matter have renewed the interest in using organic fertilizers. It is also the
only option in organic farming systems. Using organic fertilizers is also a
means of recycling nitrogen already available in the biosphere, instead of
increasing the rate of nitrogen fixation in fertilizer production. Thus,
organic fertilizers can aid in slowing down Earth’s accelerating nitrogen
cycle. However, the amount used, application time, and way to apply the
organic fertilizer can lead to inefficient use of nitrogen and carbon substrates, which promotes nitrogen loss through both NO
3 leaching and
denitrification.
7.1.2. Fertilization effects on nitrous oxide emissions
Besides promoting denitrification activity, fertilization also positively affects
the N2O emissions from agricultural soil. Higher N2O emissions in
response to fertilization could simply be due to higher denitrification rates
or an increase of the N2O/N2 ratio. By reviewing data for N2O emissions
from agricultural soils, Eichner (1990) found rates of emission ranging from
0.2 to 42 kg N2O–N ha1 year1. Calculated as the percentage of the
nitrogen fertilizer applied, nitrogen losses varied from 0.1% to 5% for
N2O (Akiyama et al., 2004; Eichner, 1990; Germon et al., 2003; Granlı́
and Bockman, 1994; Mosier et al., 1998; Sherlock et al., 2002; Whalen et al.,
2000) and 0% to 25% for dinitrogen gas (Barraclough et al., 1992; Ryden,
1983; Svensson et al., 1991). The application of 220 kg nitrogen as a mineral
fertilizer to soil induced higher N2O losses throughout the crop season
compared with an unfertilized soil (Sehy et al., 2003). In addition, Mulvaney
et al. (1997) demonstrated an increase in the mole fraction of N2O emissions
in mineral fertilized treatments compared to an unfertilized control. During
the first week of incubation, the N2O/N2 ratio was larger for ammonium
sulfate, ammonium nitrate, or mono-ammonium phosphate than for anhydrous ammonia, di-ammonium phosphate, or urea treated soil. Application
of different manures also stimulates N2O emissions and a strong effect of
poultry manure compared with swine or cattle manure was reported by
Dong et al. (2005). Accordingly, Akiyama et al. (2004) showed that emissions sewage sludge or poultry manure-fertilized soil was higher than those
from farmyard manure or composted plant residues. It has also been
demonstrated that N2O emissions increase with the amount of manure
applied (Akiyama et al., 2004; Chang et al., 1998).
The relative effect of mineral or organic fertilization on N2O emissions
is still in controversy. Ellis et al. (1998) inferred that cattle slurry application
stimulated both the total nitrogen losses and the N2O production compared
Soil Denitrifiers
277
with mineral fertilizer additions, while no difference were observed by Meng
et al. (2005). On the other hand, Lampe et al. (2006) inferred that more N2O
was emitted after mineral than cattle slurry fertilization. Different N2O/N2
ratios between organic and mineral fertilized were reported by Dittert et al.
(2005) who observed that application of either calcium nitrate or slurry
resulted in a ratios of around 1:1 and 1:14, respectively. In a long-term
field experiment, Dambreville et al. (2006a) also measured lower N2O/N2
ratio from experimental plots fertilized with pig slurry than from plots
fertilized with mineral fertilizers. When comparing organic with conventional farming practice receiving mineral fertilizers, Flessa et al.
(2002) showed that the former led to lower N2O emissions per hectare,
but yield-related emissions were the same.
An interaction between organic and mineral fertilizers was reported by
Ellis et al. (1998) who showed that N2O losses were greater following
mineral fertilizer application to soils that had previously (<5 months) been
fertilized with cattle slurry. The effect of combined organic and mineral
fertilization on increased emissions was confirmed by Dittert et al. (2005).
Application of fresh cattle slurry together with calcium nitrate increased
N2O emission six times during the first 4 days after application compared
with single application of one of the fertilizers. The easily decomposable
slurry carbon probably induced N2O emissions from the calcium nitrate
fertilizer, as indicated from 15N-labeling experiments. Similar effects were
reported by Arcara et al. (1999) when investigating additive effects of pig
slurry and urea. After comparing N2O emissions from two different soils
under different mineral nitrogen fertilization and slurry application, van
Groenigen et al. (2004) concluded N2O emissions varied with soil type,
fertilizer type, and fertilizer application rates. The importance of the soil
type was confirmed in other studies (De Klein and Van Logtestijn, 1994;
Terry and Tate III, 1980; Velthof et al., 2002; Wever et al., 2002).
Even though denitrification research in agriculture has been dealing
with the gaseous nitrogen losses for decades, there is still no clear-cut answer
to how the organic fertilizers affect the ratio of N2O to total denitrification.
Nonetheless, it can be agreed on that organic fertilizers increase denitrification. The activity of the denitrifying community is also a crucial factor in
regulating N2O emissions since denitrification is both a source and a sink for
N2O. Research is only at the beginning of resolving how the denitrifying
community is affected by fertilization and the interplay between the environmental factors, the denitrifier community structure, and nitrogen
emissions caused by denitrification.
7.1.3. Fertilization can modify denitrifier communities
How fertilization affects the sporadic events of denitrification and N2O
emissions are fairly well studied, but how the dynamics of denitrifier
population relate to fertilization practice and its importance for nitrogen
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Laurent Philippot et al.
emissions are not. Short-term effects of mineral nitrogen fertilizers on the
structure of the denitrifier community have only been investigated in a
couple of studies. Avrahami et al. (2002) explored the effect of different
ammonium concentrations (6.5, 58, or 398 mg NH4þ–N g1 of dry soil) on
the nirK genotypes using T-RFLP. After a 4-week incubation experiment,
significant differences in the nirK community structure were observed at
the two highest ammonia concentrations. In contrast, others reported that
the narG community structure was unaffected by different NO
3 concentrations after 2 weeks of incubation, although the NO
reduction
rate
3
increased with the highest amendment (300 mg NO
–N)
(Deiglmayr
3
et al., 2006).
Effect of organic fertilization on the denitrifier community composition
was investigated by a few authors in field experiments of various ages.
A 3.5-year amendment with different fermented organic fertilizer or common organic manuring practices revealed slight changes in the nirS singlestranded conformation polymorphism patterns between the treatments
(Schauss, 2006). In addition, the changes coincided with a change in the
nirS gene copy numbers, but no differences were observed in potential
denitrification rates. In a 6-year-old field trial, analysis of the denitrifier
community structure in fields treated either with mineral fertilizer (60
and 120 kg N ha1 year1) or cattle manure (75 and 150 kg N ha1
year1) showed that the main differences in nirK T-RFLP patterns were
due to seasonal variation (Wolsing and Priemé, 2004). However, small
differences that might be explained by the type of fertilizer were also
observed, whereas the amount of fertilizer did not have any effect. Similarly,
comparison of fertilization with either ammonium nitrate (162 kg N ha1
year1) or composted pig manure (213 kg N ha1 year1) during 7 years
showed significant, but small differences in structure of the narG and nosZ
communities, although the potential activity differed (Dambreville et al.,
2006b). The most complete survey on the impact of fertilization regime
on denitrifiers was performed by Enwall et al. (2005). Effects of calcium
nitrate, ammonium sulfate, cattle manure, and sewage sludge were
analyzed in an experimental field established in 1956 on narG and nosZ
communities (Fig. 5). Fingerprint analyses showed differences in the denitrifier community structure in plots treated with ammonium sulfate and
sewage sludge, which were the treatments with the lowest pH. No differences were observed between the unfertilized plots and those treated with
calcium nitrate or manure. As expected, potential denitrification rates were
higher in plots treated with organic fertilizer than in those treated with
mineral fertilization. Altogether, these results suggest that long-term fertilization can affect activity and composition of the denitrifier community
differentially.
279
Sewage sludge
Cattle manure
(NH4)2SO4
Ca(NO3)2
No fertlization
No fertlization
No crops
Soil Denitrifiers
Figure 5 Denaturing gradient-gel electrophoresis banding pattern of PCR amplified
partial nosZ genes derived from soil treated with six different fertilization regimes in triplicates at the Ultuna long-term experimental field site established in 1956 (redrawn
from Enwall et al., 2005).
8. Effect of Environmental Pollution
on Denitrifiers
8.1. Pollution affects denitrification
For agricultural soil, there is concern about responsible use and maintenance
of microbial functions and diversity for sustainable ecosystem management
and crop production. Several studies have shown that denitrification is
inhibited by organic pollutants, for example, polyaromatic hydrocarbons
(PAHs) (Richards and Knowles, 1995; Roy and Greer, 2000; Sicilano et al.,
2000) and pesticides (Bollag and Kurek, 1980; Pell et al., 1998), in addition
to heavy metals (Bardgett et al., 1994; Bollag and Barabasz, 1979; HoltanHartwig et al., 2002; McKenney and Vriesacker, 1985). It is also known that
the enzymes involved in the denitrification chain are differently affected by
various stress factors, with N2O reductase being the most sensitive (Bonin
et al., 1989; Firestone et al., 1980; Holtan–Hartwig et al., 2002; Sicilano
et al., 2000). Inhibition of this enzyme results in increased production of
N2O, and this has been shown to be the case in heavy metal contaminated
soil (Vásquez-Murrieta et al., 2006). PAHs are not considered a big problem
in agroecosystems, although they can reach agricultural soil accidentally by
280
Laurent Philippot et al.
deposition. The same goes for heavy metals. However, copper is still used as
a fungicide, mainly in organic farming systems and vineyards. In addition,
some organic fertilizers from different waste residues may be contaminated
by heavy metals. Pesticides, on the other hand, are frequently used and are
often crucial for reaching sufficient crop yields in conventional farming
systems. Similarly to fertilizers, the use of pesticides is expected to increase
globally during the next 50 years by nearly three times reaching 107 metric
tons year1 (Tilman et al., 2001). Therefore, more analyses of the response
of denitrifiers after the application of pesticides to agricultural soil are highly
warranted.
8.2. Pesticides
The influence of pesticides on different nitrogen transformation processes
has mainly been studied for nitrification (Gadkari, 1988; Sattar and
Morshed, 1989; Stratton, 1990). In most cases, nitrification rates were
significantly reduced. This could be explained by the nearly monophyletic
nature of the ammonia-oxidizing bacteria, associated to the first step of
nitrification. Even if the chrenarchaeal nitrifiers are taken into account, the
taxonomic diversity, as we know it today, is rather limited. In contrast, the
reaction patterns of denitrifiers in response to pesticide application are not as
clear, which may be related to the vast number of taxonomically, distantly
related genera of denitrifiers in soils.
8.2.1. Inhibition or stimulation of denitrification activity
Early studies by Bollag and coworkers (Bollag and Kurek, 1980; Bollag and
Nash, 1974) reported an accumulation of NO
2 and N2O in soils incubated
under anaerobic conditions when derivatives of the insecticide chlordimeform ([N-4-chloro-o-tolyl]-N 0 ,N 0 -dimethylformamidine) were added.
Interestingly, this inhibition was not caused by the insecticide itself but by
the metabolites formed during degradation (N-formyl-4-chloro-o-toluidine
and 4-chloro-o-toluidine). The same researchers also showed that aniline
intermediates of other pesticides have stronger inhibitory effects on denitrification in soil than their parent compound. The most comprehensive study
on the effect of pesticides on denitrification was conducted by Pell et al.
(1998). The acute toxic effect of 39 herbicides, 10 fungicides, and 5 insecticides was tested on potential denitrification activity in one Swedish agricultural soil. They demonstrated that 23% of the pesticides tested at 100 mg
active ingredient g1 dry soil had an effect on potential denitrification
(Table 2). For example, potential denitrification was stimulated by the
addition of AMPA or fenvalerate, whereas the herbicide ioxynil, the fungicides mancozeb and maneb, and the insecticide zineb showed the most
pronounced inhibition of denitrification activity. Other studies also reported
an inhibitory effect of maneb (Bollag and Henninger, 1976) and mancozeb
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Soil Denitrifiers
Table 2 Effect of 54 pesticides and their degradation products on the Potential
Denitrification Activity (PDA)
Pesticide
Herbicides
Alloxydim
AMPA
Atrazine
Bentazone
Bromacil
Chlorbromuron
Dalapon-Na
2,4-D
2,4-DB
2,4-DP
2,4-Dichlorophenol
Diuron
Glyphosate free acid
Glyphosate isopropylamid
Hexazinone
Imazapyr
Ioxynil
Lenacil
Linuron
MCPA
MCPB
MCPP
Metobromuron
Metribuzin
Napropamide
p-Chlorophenoxy
acid
Picloram
Propazine
Effect on PDA
(% of control)
115 5
135 8*
98 5
97 5
101 3
102 2
119 16
112 3*
115 2*
115 9
75 1*
101 5
115 7
97 7
106 2
n.a.
31 8*
97 2
113 1
114 6
107 4
99 6
103 8
86 4*
101 1
101 2
117 3
114 2
Pesticide
Effect on PDA
(% of control)
Herbicides
Simazine
TBA
TCA
2,4,5-T
Terbuthylazine
Tertbutryn
Tri-allat
Triclopyr
Trifluralin
97 13
96 6
114 5
104 3
99 0
93 3
97 7
n.a.
96 2
Fungicides
Benomyl
Carbendazim
Iprodione
110 6
92 3*
113 11
6 1*
7 4*
110 8
Mancozeb
Maneb
Thiphanatemethyl
Triadimefon
Triadimenol
Vinclozolin
Zineb
72 5*
99 5
122 10
33 1*
Insecticides
Aldrin
Cyromazine
Fenvalerate
96 6
97 1
121 6*
Heptachlor
Permethrin
94 4
101 6
Figures given are mean values standard deviation (n ¼ 3) in percentage of a control soil without
pesticide addition (from Pell et al., 1998).
* significantly different from control in Student’s t-test (p< 0.05).
(Kinney et al., 2005) on denitrification. In the latter, the authors observed
inhibitory effects of the fungicides mancozeb and chlorothalonil, and the
herbicide prosulfuron on denitrification with increasing pesticide concentration, ranging from 0.02 to 10 times that of a standard application rate.
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Laurent Philippot et al.
The absence of significant effects of dalapon, atrazine, and simazine on denitrification in soil observed by Pell et al. (1998) was also reported by Yeomans
and Bremner (1987) at concentration ranging from 5 to 100 mg g1 soil.
Contradictory results on the effect of the same pesticides were also
reported. Yeomans and Bremner (1985a,b) tested the effect of 20 herbicides,
7 insecticides, and 6 fungicides on denitrification in different soils. At a
concentration of 10 mg g1 soil, none of the pesticides tested affected denitrification. When applied at a concentration of 50 mg g1 soil, only the fungicide
captan inhibited denitrification while mancozeb or maneb, had no effect or
enhanced denitrification. The herbicides butylate, EPTC, diuron, simazine,
and dalapon had no effect on denitrification and all the others either enhanced
or inhibited denitrification, with effects varying according to the soil type.
These contradictory results could be at least partially explained by the work of
Tu (1994). They showed that the majority of 14 insecticides applied to a sandy
soil had a significant effect on denitrification during the first week after
application, but most of the effect had disappeared after 2 weeks of incubation.
These results suggest that denitrification activity has a high capacity to return to
its initial level after a temporary disturbance.
To summarize findings in the literature, fungicides have more often
been reported to have a negative impact on denitrification activity than
herbicides. The impact of pesticides on denitrification activity in soil is
likely to be dependent on the soil type, the concentration and nature (pure
active ingredient or formulated preparation) of the pesticide applied, the
climatic conditions, and in which way it is degraded. Addition of pesticides
can reduce bacterial denitrification, probably due to cell death or cell
inactivation, but it can also stimulate this process due to (1) the use of the
pesticide as an electron donor by the denitrifiers; (2) death of organisms
caused by the pesticide, which results in an easily available source of carbon
for denitrification; or (3) an unspecific stress response. However, stimulation of denitrification in response to pesticide addition is a symptom that
should be considered just as severe as decrease of denitrification.
8.2.2. Pesticide effects on denitrifier community structure
Only a few studies investigated the effect of pesticides on the size and the
structure of the denitrifier community. Cen et al. (2002) investigated effects
of carbofuran, carbendazim, and butachlor during 4 weeks on the population size of denitrifying bacteria and their activity in different Chinese paddy
soils. Lower concentrations of the pesticides (1 mg g1 dry weight soil) in
general increased the population size and activity, whereas higher concentrations reduced both parameters. Increased numbers of denitrifiers were
also reported after addition of 50–300 mg g1 soil of malathion (GonzálezLópez et al., 1993) and 5–10 kg ha1 of captan or alachlor (Martı́nezToledo et al., 1998; Pozo et al., 1994). Similarly, addition of the herbicide
Topogard 50 WP at a concentration of 3 kg ha1 in soil with varying pH
Soil Denitrifiers
283
resulted in a temporary increase of denitrifiers, and the effect of Topogard
on bacteria was likely to be dependent on soil pH (Kara et al., 2004). In a
study, Philippot et al. (2006) analyzed the community structure and the
activity of the NO
3 -reducing bacteria in a maize field treated with atrazine
or glyphosate. While temporal shifts in both structure and activity of the
denitrifying community were recorded after 8 years of cultivation, no
pesticide effect was observed.
8.3. Heavy metals
8.3.1. Negative effects on denitrification activity
In contrast to organic pesticides, denitrifiers are highly sensitive toward
heavy metal stress. Denitrification has been shown to be more sensitive to
heavy metals than aerobic soil respiration (Bardgett et al., 1994). Also,
Kandeler et al. (1996) concluded that heavy metals influenced enzymatic
processes in nitrogen cycling more negatively than those in carbon cycling.
Hence, denitrification tests have been used to assess the presence of bioavailable heavy metals in soil (Speir et al., 2002). Heavy metals such as arsenic,
cadmium, chromium, copper, lead, silver, and zinc have all shown a negative
effect on denitrification activity in soil and sediment, and the effect is usually
immediate (Bardgett et al., 1994; Bollag and Barabasz, 1979; Holtan–
Hartwig et al., 2002; Johansson et al., 1998; McKenney and Vriesacker,
1985; Probanza et al., 1996; Sakadevan et al., 1999; Throbäck et al., 2007).
Interestingly, Holtan-Hartwig et al. (2002a) observed that the N2O
reduction was more affected than the N2O production rate after addition
of a heavy metal mixture of Cd, Cu, and Zn at different concentrations.
After incubating the soil for 2 months, a complete recovery in denitrification activity and N2O production rate was shown, but the N2O reduction
capacity was still not fully restored. Also, Bollag and Barabasz (1979)
observed an increased accumulation of N2O from soil incubated with Cd,
Cu, Zn, and Pb. These findings indicate a more severe inactivation of the
N2O reductase by heavy metals than other enzymes in the denitrification
cascade. Differences in resistance to heavy metals among soil denitrifier
communities are probably large and depend also on soil chemical properties,
such as pH, cation exchange capacity, and organic matter content, which
determine the bioavailability of metals. Reduced availability of the heavy
metals is expected in clay soils those with high organic matter content.
8.3.2. Heavy metal effects on community composition and
abundance of denitrifiers
The most commonly reported effect of heavy metals on microbial communities is decreased genetic diversity (Kozdrój and Van Elsas, 2001; Moffett
et al., 2003; Muller et al., 2002). Nevertheless, increased bacterial diversity
has been observed in soil after 10–60 days of Cu, Cd, and Hg exposure
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(Ranjard et al., 2006), and after long-term applications of sewage sludge
with high levels of Cu, Ni, Cd, Zn, and Cr (Sandaa et al., 2001). Giller et al.
(1998) explained increased diversity in heavy metal contaminated soils with
the intermediate disturbance hypothesis. This hypothesis postulates that stable
environments with high numbers of competitive species have increased diversity because metal stress reduces the innate competitive exclusion between
bacterial populations and induces enrichment of other populations. The
diversity continues to increase until eventually the stress becomes so high
that it begins to decrease.
How the diversity or community composition of soil denitrifiers is
affected by heavy metal pollution is not well studied. Holtan-Hartwig
et al. (2002a) showed that soil-extracted denitrifier cells exposed to heavy
metals developed a higher tolerance to these after 2 months. As the community metal tolerance progressed, estimated growth rates were lowered. In
soil spiked with silver, the kinetically derived specific growth rate of the
denitrifying community indicated that part of the community was resistant
to silver, although there was a negative impact on soil microbial biomass
( Johansson et al., 1998). In a follow-up study, it was shown that silver
increased the diversity of denitrifiers in soil and induced enrichment of a
certain clade of nirK denitrifiers (Throbäck et al., 2007). However, the
number of nirK-type denitrifiers was negatively correlated with increasing
concentrations of silver. The specific activity (k0), determined as the potential denitrification activity per nirK copy number, was also shown to
decrease with increasing silver concentrations, which indicates that physiological properties of the denitrifiers could be affected by heavy metals
(Throbäck et al., 2007). A detailed analysis of heavy metals effects on
Proteobacteria was done with a system-biology-like approach by Kesseru
et al. (2002), using Ochrobactrum anthropi a well-known Gram-negative
bacterium as a model. Surprisingly, the cells were able to denitrify even in
the presence of high concentrations of different heavy metals. The reason
for that might be the good nutrient status in the media, which gave the
organisms enough energy to protect themselves against the heavy metals.
Bacterial communities can develop heavy metal tolerance (Bååth, 1989;
Mergeay et al., 2003) and within the a-, b-, and g-subdivison of Proteobacteria, where many denitrifiers belong, it is known that several genera show
high tolerance to heavy metals. There are several reasons for increased
tolerance, such as substitution of sensitive strains by tolerant ones, spread
of resistance genes, and genetic modifications to produce heavy metal
resistance. The transfer of genes coding for resistance or tolerance against
heavy metals by plasmids among bacteria of different phylogenetic groups in
soil has been described in many studies not exclusively related to denitrification (Smalla and Sobecky, 2002). We should be aware of that increased
heavy metal resistance often is connected to antibiotic resistance. The
emergence of community resistance against heavy metals, or other types
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285
of pollutants, can be regarded as a process describing deterioration of the
ecosystems. We should therefore continue with risk assessment of pollutants
in agricultural soils.
9. Conclusions and Outlook
Agronomical practices resulting in an increase of carbon or NO
3
availability or a decrease of the oxygen partial pressure stimulate the denitrification activity in soil, which can cause major losses of nitrogen and
emissions of N2O. The denitrification activity is rapidly regulated by these
factors and therefore, the effect of agricultural practices on denitrification
activity can be observed very fast. For example, NO
3 fertilization results in
higher nitrogen gas fluxes by denitrification within few days or even hours if
the oxygen partial pressure is low in the soil, for example, after a heavy
rainfall. The factors regulating denitrification interact, which makes it
difficult to interpret the highly variable effects measured in the field. Future
research should consider the small-scale heterogeneity, including soil aggregates and other hotspots in soil to deepen our understanding of the regulation of denitrification. Experimental studies have pointed out the necessity
of taking into account such microheterogeneities (Sexstone et al., 1985;
Sierra and Renault, 1996) and the microscale approach to study denitrification is motivated by the fact that conditions experienced by soil organisms
are not reflected by measurements of these conditions made on bulk soil
samples (Parkin, 1987). For example, O2 concentrations may decrease from
values nearly equal to the atmospheric concentration to zero values within a
few millimeters in soil clods (Curie, 1961; Sexstone et al., 1986; Sierra et al.,
1995). For readily decomposable organic matter particles, a thin layer of
covering water with a thickness of about 160 mm may be sufficient for
anaerobiosis to occur (Parkin, 1987). On the other hand, it is important to
upscale and generalize the results from 1 g of soil to the field or landscape
scale to make them appealing for model developers and decision makers.
This requires integration of geostatistical tools and application of remotesensing techniques to identify landscape patterns. Another question, that we
have to answer in the future if we want to integrate the data on larger scales
is, if we have not overseen important hotspots for denitrification. For
example, it is known that denitrifiers in the earthworm gut is involved in
the in vivo emission of N2O by earthworms and denitrification also occurs in
earthworm casts (Horn et al., 2006).
In contrast to the weak resistance to changes in environmental conditions observed for denitrification activity, summarizing the data available on
how the total denitrifier community composition responds to various
agronomical practices suggests that it exhibits a high capacity to withstand
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Laurent Philippot et al.
perturbations. Hence, major modifications of the community structure
were only observed in long-term experiments by which the soil’s physical
and chemical parameters were also modified, whereas most of the laboratory
experiments lasting only weeks or months resulted in minor or no modifications. Even though carbon and NO
3 strongly affects the activity, these
factors apparently do not drive the composition of this functional community in agricultural soil. Long-term amendment of organic fertilizer or
addition of root-derived carbon only slightly affected the denitrifier community structure. Similarly, addition of NO
3 at high concentration in a
3-week laboratory experiment or at 80 kg N ha1 year1 during 50 years in
the field did not result in significant differences compared with the controls.
The understanding of mechanism connecting denitrifier diversity and activity has mainly been based on some nice data sets of the genetic potential
using DNA analysis and potential denitrification rates. These results also
emphasize the redundancy of functional genes involved in denitrification.
However, we do not know if a change in the diversity or composition of the
denitrifier community plays a role for denitrification activity or N2O fluxes.
Since the denitrifier communities represented by the total gene pool seem
to be highly resistant to changes, a better understanding might be gained by
focusing on the active denitrifiers under different conditions. More data on
the induction of a genetic potential by targeting gene transcripts and the
active enzymes in denitrification is needed if actual denitrification rates
occurring in the field are to be explained.
Due to the great taxonomic and physiological diversity of denitrifiers,
processes shaping the denitrifier community structure are probably not
different from those shaping other heterotrophic bacterial communities.
Thus, bacterial DNA-fingerprinting analysis of 98 soil samples from across
North and South America revealed that soil pH was the best predictor of
bacterial community composition (Fierer and Jackson, 2006). Similarly,
several authors investigating effects of fertilization regimes or increase of
CO2 concentration reported that the main differences in the structure of the
denitrifier community were linked to soil pH rather than the treatment per
se, which again implies that pH is an important driver of the denitrifier
community composition. Since the diversity of the denitrifiers is governed
by factors that also shape the diversity of other heterotrophic bacteria and
the fact that denitrifiers are facultative, the diversity of the denitrifier
community can be affected by agricultural practices independently from
the denitrification trait of its members. While no effect of such agricultural
practices will be observed on denitrification on a short-term basis, it can
result in changes in the denitrification activity on a long-term basis. This is
true if the populations in the modified denitrifier community exhibit different physiological properties and react in different manners to additional
perturbations. It is therefore important to consider the consequences of
agronomical practices in a dynamic manner. Investigating consequences
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287
of agronomical practices under a limited number of environmental conditions is not very helpful and informative to predict the effects of a modified
denitrifier community on the functioning of this process. Unfortunately, it
is not conceivable to compare the effect of agronomical practices on
denitrification activity under all possible environmental conditions or stresses, which can be faced today or in the future. Learning more about the
ecology of denitrifiers, integrating structure and function of this community
in soil, and developing methods to do that is essential for answering questions concerning nitrogen economizing and environmental impact from
modern agriculture. It is also in line with the main challenge in microbial
ecology—the understanding of the role of biodiversity for ecosystem functioning. In contrast to plant and animal ecology, which emerged in the
nineteenth century and generated most of the general ecological theories,
microbial ecology is a relatively young science and ecological theories and
concepts are still under construction.
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