Ecology of Denitrifying Prokaryotes in Agricultural Soil

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

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2007 Elsevier Inc. All rights reserved.

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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.

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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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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.

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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 ha
1 can be lost through denitrification during 1 year under extreme conditions, although values in the range 0–200 kg N ha
1 year
1 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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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 year
1) of the total annual global emission with about 6.3 Tg year
1 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 liter
1 in drinking water for human consumption.

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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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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) (Vo¨lkl 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) (Che`neby 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) (Che`neby 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)

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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) (Che`neby 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)

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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).

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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 (Che`neby 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

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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.

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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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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.

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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 (Ro¨ver 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, Do¨rsch 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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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; Mu¨ller et al., 2002) and lysis of microorganisms (Schimel and Clein, 1996; Skogland et al., 1988). Mu¨ller 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

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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 No¨mmik, 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 (Do¨rsch and Bakken, 2004; HoltanHartwig et al., 2002; Melin and No¨mmik, 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. Prieme´ 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 m
2, and between 13 and 340 mg N–N2O m
2 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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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

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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 kg
1 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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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 (Hu¨tsch 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

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performed using the most-probable-number method with NO
3 agar broth as growth medium revealed densities ranging between 103 and 106 bacteria g
1 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 (Che`neby et al., 2004). Density of denitrifiers was 2.4  106 and 1.6  107 cells g
1 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 ` neby of NO
3 -reducing actinobacteria by the maize roots. In contrast, Che 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

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(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 (Hu¨tsch 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

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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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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

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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; No¨mmik, 1956; Sˇimek 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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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. (Sˇimek and Hopkins, 1999; Sˇimek 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 Prieme´, 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

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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 (Leve´n 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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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 ha
1 year
1. 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; Granlı´ 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

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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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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 g
1 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 ha
1 year
1) or cattle manure (75 and 150 kg N ha
1 year
1) showed that the main differences in nirK T-RFLP patterns were due to seasonal variation (Wolsing and Prieme´, 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 ha
1 year
1) or composted pig manure (213 kg N ha
1 year
1) 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 (Va´squez-Murrieta et al., 2006). PAHs are not considered a big problem in agroecosystems, although they can reach agricultural soil accidentally by

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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 year
1 (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 g
1 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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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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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 g
1 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 g
1 soil, none of the pesticides tested affected denitrification. When applied at a concentration of 50 mg g
1 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 g
1 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 g
1 soil of malathion (Gonza´lezLo´pez et al., 1993) and 5–10 kg ha
1 of captan or alachlor (Martı´nezToledo et al., 1998; Pozo et al., 1994). Similarly, addition of the herbicide Topogard 50 WP at a concentration of 3 kg ha
1 in soil with varying pH

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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; Throba¨ck 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 (Kozdro´j 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 (Throba¨ck 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 (Throba¨ck 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 (Ba˚a˚th, 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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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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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 ha
1 year
1 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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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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