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While available organic C and oxygen have a major impact on total denitrification, soil pH influences mainly the ratio of N2O/N2 (Šimek and Cooper, 2002). Maximum denitrification occurs usually at highest pH values and a linear relationship is often observed between dentrification and pH (Focht, 1974). There is a tendency that the N2O/N2 ratio increases with decreasing pH (Burford and Bremner, 1975; Liu et al., 2010; Šimek and Cooper, 2002) and a more complete reduction towards N2 is found at higher pH values (Nõmmik, 1956). The effect of pH may be due to an impact on substrate availability which influences membrane permeability or speciation of chemical species (Dassonville and Renault, 2002). N2O reductase is more sensitive to pH than other reductases in the denitrification sequence and therefore low pH may increase the lag phase of de novo synthesis of this enzyme (Liu et al., 2010; Šimek and Cooper, 2002). Apparently the detrimental effect of low pH on N2O reductase occurs at the post-transcriptional level by interfering with translation, protein assembly or an effect on the activity of functional enzymes (Liu et al., 2010). Microsite variations in soil pH may support a diverse microbial community which exhibits a different sensitivity to pH (Cavigelli and Robertson, 2000; Šimek and Cooper, 2002). Also chemodenitrification may play a role for N2O and N2 production in acidic microsites (Chalk and Smith, 1983; Stevens et al., 1998).
Microbial community dynamics Microbial community dynamics are influenced by the interacting effect of environmental conditions together with the substrate relationships (carbon substrates) and the availability of electron acceptors (N-oxides). The most important environmental regulators for denitrification are: oxygen pH temperature (1974). Mineralisation activity provides substrates and is closely related to moisture/oxygen-temperature conditions (Focht and Verstraete, 1977; Morley and Baggs, 2010;
Myrold and Tiedje, 1985b). For instance a high rate of soil organic C content mineralisation favours O2 consumption. In fertilized soils, C rather than NO3- availability limits total denitrification, but NO3- has a pronounced effect on the ratio of N2O/N2 (Beauchamp et al., 1989; Dendooven and Anderson, 1994; McCarty and Bremner, 1993; Myrold and Tiedje, 1985a, b; Nõmmik, 1956). Many studies have focused on either proximal or distal factors, as described by Groffman (Groffman et al., 1988), but only a few studies so far have tried to link these factors. The C substrate determines the efficiency with which N oxides (NO3-, NO2-) are reduced (Beauchamp et al., 1989). Denitrifiers are able to compete successfully for C with other heterotorophs (Myrold and Tiedje, 1985b). Thus
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specific microbe-substrate relationships exist that may explain the link between decomposition products and microbial populations (1989). However, these relationships are still not well understood. The availability and identity of the C substrates and in particular the ratio between C substrate to e--acceptor also governs whether denitrification or DNRA occurs in soils. While the need for C is universally acknowledged there is still a need to identify assays that can predict forms of C substrate utilised by various microbial groups (Rütting et al., 2011).
Successive steps in the reductive denitrification sequence exhibit differing sensitivities to oxygen (Betlach and Tiedje, 1981). Higher N2O emissions and increasing N2O/N2 ratios are observed with increasing aerobicity (Focht, 1974; Payne, 1973). Aerobic and anaerobic microsites, as indicated by different redox potentials, co-exist in soil may support different microbial communities. To advance the mechanistic understanding of microbial N cycling in terrestrial ecosystems there is a need to link microsite and not simply bulk soil conditions to microbial activities.
In natural environments diverse microbial populations are observed (Tiedje, 1988). The greatest spatial aggregation is generally observed in the top soil (Nunan et al., 2002). Spatial distributions may develop to the patchy distribution of organic material (Parkin, 1987) and in response to environmental conditions (Tiedje, 1988) such as aerobic-anaerobic microsites. There is still a need to better understand the microenvironment in where N transformations occur (McLaren, 1970).
Microbial processes are affected by substrate availability, thermodynamic regulation and inhibition, pH, and redox potential (Brock, 1967; Dassonville and Renault, 2002; Firestone and Davidson, 1989). Interactions between these factors are only poorly understood. Cavigelli and Robertson (Cavigelli and Robertson, 2000) investigated for instance the effect of pH and oxygen on the N2O/N2 ratio in both an arable and a non-disturbed successional field and found that the observed differences were most likely governed by the prevailing microbial community.
New conceptual framework for N dynamics in soil There is a need for new theoretical and conceptual frameworks to understand how microorganisms influence ecosystem processes (2011). Two aspects need to be taken into account to understand the relationships between environmental regulators and microbial activity. First, there needs to be an acknowledgement that different microsites exist in the soil, which to date have been considered as homogenous spaces with respect to environmental regulators. Secondly, there is a requirement to identify the units that represent physiologically similar microbial entities that live in mutualistic or antagonistic relationships within these microsite.
For instance while anaerobic conditions are required for denitrification to procede, aerobic denitrification (NO3- reduction under microaerophyillic conditions) has been observed (Bréal, 1892). The co-respiration probably occurs under conditions when respiratory pathways are rate limited by oxygen and additional NO3- reduction would allow faster growth rates (Robertson and Kuenen, 1984, 1990). Furthermore, aerobic denitrification appears to be linked to heterotrophic nitrification, and is possibly a detoxifying mechanism to reduce high NO2- concentrations (Robertson and Kuenen, 1990) converted to gaseous N forms such as nitrous oxide (N2O).
Nitrous oxide is predominantly produced in soils and emitted to our atmosphere. Thus, to understand N2O production from terrestrial ecosystems it is essential to understand the potential production processes. Apart from nitrifiers und denitrifiers, N2O is also produced by nitrifiers
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which are paradoxically denitrifying via a process called nitrifier-denitrification (Wrage et al., 2001). Nitrifier denitrification appeared to be much less important than classical nitrate-driven denitrification. But Wrage et al. (Kool et al., 2011) showed that this pathway can be a major source for N2O if conditions for denitrification are not optimal. Thus this pathway should not be ignored as a source of N2O from soil (Kool et al., 2011; Wrage et al., 2001).
Furthermore chemical transformations may also affect the N cycle transformations and losses.
Chemodenitrification may occur in soils, including the van Slyke reaction where organic N together with NO2- reacts to N2 (Tiedje, 1988). However, our understanding of the potential contribution of chemodenitrification to gaseous N losses as nitric oxide (NO) and N2O is far from complete as demonstrated by recent studies and further studies on chemodenitrification are warranted, especially under high N input agricultural systems (Spott et al., 2011). Besides bacteria, fungi are capable N transformations such as NO3- assimilation and denitrification when NH4+ is depleted (Laughlin and Stevens, 2002; Zumft, 1997). However, there is a dearth of information on fungal denitrification (Morozkina and Kurakov, 2007).
2. Modelling concepts Various kinetics expressions have been used to model the dynamics of the various N oxides in the denitrification sequence (Kohl et al., 1976) including zero-order (Focht, 1974) first-order (Bowman and Focht, 1974) or Michaelis-Menten kinetics (1981). Betlach and Tiedje (Betlach and Tiedje,
1981) concluded that kinetic models can be a useful guide to evaluate the physiology of denitrification. Most denitrification models now consider a Michaelis-Menten type formulation for each enzymatic step and typically may consider dual substrate reactions (Betlach and Tiedje, 1981;
Bowman and Focht, 1974; Focht, 1974). Simple kinetic expressions are restricted to situations where no competition between individual N oxides takes place. Therefore, models based on competitive kinetics have been proposed that allow a branching of the electron flow from a electron donor towards various N oxides (Almeida et al., 1997; Cho and Mills, 1979; Thomsen et al., 1994).
In these models each reduction step is governed by the overall availability of electron donors or denitrification intensity and the availability of N oxides for each reduction step (Dendooven et al., 1994). Furthermore, competition between organisms for NO3- (e.g. denitrifier, dissimilatory reduction of NO3- to NH4+ ) is predominantly governed by the carbon availability. Thus key for the successful evaluation/prediction of heterotrophic activity in aggregated soil, is a detailed knowledge of the kinetic parameter profiles of the individual microbial groups (Tiedje et al., 1982).
Diffusion and conditions in aggregated soil Kinetic parameters in denitrification models that are applied to field situations are most likely not representative of the true kinetic reaction at the reaction sites because heterogenic distribution and diffusion of N oxides and electron donors has to be taken into account (Betlach and Tiedje, 1981;
Focht, 1974). In aggregated soils it is important to consider diffusion to the centre of aggregates where denitrification predominantly occurs (a kind of “hot spot” for denitrification) (Arah, 1990b).
Therefore mechanistic models for denitrification need to include microsite specific conditions and microbial interactions (Arah, 1990b; Myrold and Tiedje, 1985a). In particular the distribution of denitrifiers and their activity should be included across the range of soil aggregates (Miller et al., 2009). In diffusion-reaction models which take into account the aeration status, the diffusion of oxygen (Greenwood and Goodman, 1964; McElwain, 1978) and of the various N species are
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required to predict denitrification in aggregated structures (Arah, 1990a; Burford and Stefanson, 1973). These models usually include diffusion-reaction notations with microbial N uptake based on Michaelis-Menten kinetics (Greenwood and Goodman, 1964).
General model development Models for microbial N dynamics should be embedded in larger models that predict microbial process and their link to nutrient cycling and take into account the high spatial and temporal nature of the processes (Groffman, 1991). Not only a single relationship but the complexity of microbial reactions that reflect the high spatial and temporal variability needs to be considered to adequately connect ecological theory with molecular-microbial composition and function (Groffman, 1991).
Thus spatial and temporal explicit relationships between the e.g. denitrifier community and environmental regulators should be taken into account. Furthermore, a bottom-up approach should be considered where spatial and temporal variability of microbial dynamics at the microsite level will be used for instance to predict denitrification at the field and regional scale. Furthermore, as mentioned above, e.g. denitrification models should be embedded in models that simulate the C-N interactions in soil (e.g. DNDC,Li et al., 1992; NCSOIL, Molina et al., 1983; CENTURY, Parton et al., 1987). Currently no model exists that takes into account all these aspects, in particular the high spatial and temporal nature of microbial processes.
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