14
Some would use a term such as pollutant when considering the occurrence of nitrate in soil whereas others may prefer to describe soil nitrate as an essential nutrient for ecosystem sustenance. One group would seek to minimize levels of the alleged environmentally offensive substance, whereas those interested in plant biomass production would seek to optimize nitrate production. Both groups would underscore the necessity of understanding processes leading to losses of soil nitrate to achieve their objectives.
To a great degree, the existence of such divergent opinions regarding the value of nitrate in soil results from the propensity of this compound to be in the wrong place at the wrong time at undesirable concentrations. Nitrate ions in the root zone of a food crop during maximal biomass production are inarguably beneficial whereas migration of that same nitrate below the root zone to underlying aquifers transforms a beneficial soil chemical component into an undesirable water contaminant.
Societal implications associated with the timing and location of accumulation of significant soil nitrate loadings reach far beyond this conflict of development of a soil nitrate balance sheet and agreement on the best nitrate management plans for maintenance of an ecosystem. For example, traditional considerations have included concerns that nitrate contents of soil leachates or runoff waters may alter regional water quality (e.g. augmented lake eutrophication rates) and that there are potential public health difficulties (methemoglobenemia) associated with elevated nitrate loadings in drinking waters. Other concerns include management of fixed nitrogen amendment to soil systems to minimize societal costs and unwanted environmental impacts. Excessive losses of fertilizer nitrogen, whether added to agricultural, urban, or reclaimed soils, alter the economics of biomass production and increase the energy expenditures associated with industrial nitrogen fixation.
Furthermore, increased utilization of industrially fixed nitrogen may have an environmental impact far beyond the localized site. Fertilizer production requires large quantities of fossil fuels. Utilization of these resources to produce plant nutrients not only decreases reserves of a nonrenewable commodity but also increases greenhouse gas production through its conversion to carbon dioxide and energy. (See van Groenigen et al. 2015 and Denk 2017 for more detailed presentations of the current status of our understanding of these processes.)
Denitrification defined: If nitrogen fixation is used as a reference for the initiation point of the nitrogen cycle, denitrification is its termination. Denitrification is the biological reduction of nitrogen oxides to nitrous oxide and/or dinitrogen. The process is catalyzed by a diverse group of microbes, including denitrifying and anammox bacteria and fungi that use nitrous oxide, nitrate, or nitrite as terminal electron acceptors in their respiratory processes associated with energy generation (see Long et al. 2013 and Strohm et al. 2007 for more detailed introductions to the growth and energy transformations involved). Since the electrons are more readily transferred to molecular oxygen compared to nitrogen oxides, denitrification can only occur under anoxic conditions. From this basic definition of denitrification, the complexity of controlling its occurrence in heterogeneous soils becomes evident. For denitrification to occur, the microorganisms must have an energy supply (usually an organic carbon source) and reside in an anaerobic soil microsite yet most soil systems are predominantly aerobic and the primary processes producing nitrate also have an obligatory requirement for molecular oxygen. That is, within the soil system itself, the primary source of the initial substance in the denitrification pathway, nitrate, is autotrophic nitrification, an obligatorily aerobic process.
These limited examples of the role of denitrification in controlling levels of nitrate occurring in soil and the availability of this primary plant nutrient within soil ecosystems underscore the need to develop an understanding of the basic properties of the process and the factors in the soil environment that control the rate of reduction of nitrate to dinitrogen. Thus, this chapter has the overall objective of elucidating the basic biological and biochemical properties of denitrification and determining the extent of its occurrence in terrestrial ecosystems, especially in intensively cultivated agricultural soils but also in wetland sites that are or could potentially bear the brunt of the influx of water from associated agricultural soils (e.g. Burgin et al. 2010; Rousk et al. 2018).
14.1 Pathways for Biological Reduction of Soil Nitrate
Soil fixed nitrogen resources may be conserved through both assimilatory and dissimilatory nitrate reductive processes. They may also be lost from the soil ecosystem by dissimilatory nitrate reduction. In this context, “conserved” refers to retention within the soil ecosystem (generally as biomass, organic matter, or ammonium).
Assimilatory and dissimilatory nitrate reduction both involve the transfer of electrons to nitrogen oxides but they differ in the ultimate fate of the reduced nitrogen atom. With assimilatory reduction, nitrate is reduced to ammonium, which is subsequently incorporated into cellular biomass. Thus, with this process, the quantity of nitrogen reduced is proportional to cellular requirements for biomass production. In contrast, for dissimilatory nitrate reduction, the nitrogenous compounds are accepting electrons in support of cellular respiration. The final products (dinitrogen, nitrous oxide, or ammonium) are released from the cell and accumulate in the environment in concentrations far beyond those necessary for biomass synthesis. Note from the view of retention of fixed nitrogen within the soil ecosystem, dissimilatory processes can lead to losses (production of dinitrogen and nitrous oxide) or conservation (production of ammonium). These two fates of the nitrate reductive processes (dissimilatory vs assimilatory reduction) are characterized by commonalties of some of the intermediates and products (e.g. ammonium, nitrate). They are also distinguished by differences in the specific enzymes catalyzing the reductive reactions.
Since assimilatory reduction of nitrate is closely linked to biomass production, it is catalyzed by all soil inhabitants capable of using nitrate as a nitrogen source. Ammonium, the terminal product of the process, is incorporated into plant and microbial biomass. Also, although a portion of the cellular reducing power is transferred to the nitrogen, the cells rely upon other pathways as terminal electron acceptors (e.g. molecular oxygen, fermentative products).
Three commonly evaluated microbial processes are classed under the heading of dissimilatory nitrate reduction. These are distinguished by their respective products: (i) nitrite, (ii) ammonium, and (iii) nitrous oxide and dinitrogen (denitrification).
The first of these dissimilatory reductions involves the simple reduction of nitrate to nitrite. The general ability of many bacterial species to catalyze the process has led to the use of nitrate reduction to nitrite as a reliable taxonomic test for grouping bacterial isolates into genera and species. This reductive process is also significant in some in situ soil systems. In a soil site overlain by oxygen‐bearing water, a nitrite and nitrate cycle that provides energy for nitrite oxidizers (see Figure 11.3) may develop. Under anoxic conditions, nitrate is reduced to nitrite by portions of the general soil bacterial population. This nitrite then diffuses from its source in the anoxic microsites to aerobic microsites where the autotrophic nitrite oxidizers utilize it in their energy‐generating processes. Evidence for the occurrence of such interactions in native sites is the occurrence of an imbalance between ammonium and nitrite oxidizer populations in some soil ecosystems. The nitrite oxidizer populations are larger than would be expected to occur were ammonium oxidizers the sole provider of their energy resource, nitrite.
The second and third members of our list of dissimilatory nitrate reduction processes, dissimilatory ammonium production and denitrification, have several common characteristics. Both processes are coupled to organic matter decomposition (i.e. oxidation of fixed carbon compounds provides cellular energy and reducing power), the nitrogen oxides that are reduced serve as terminal electron acceptors, and growth yields of the bacteria are increased as a result of the energy provided by passage of an electron transport pathway that includes cytochromes (see Section 14.2 for discussion of the role of cytochromes in denitrification.).
Although not previously considered to be a major soil process, ammonium production from nitrate reduction has been shown to occur in soils and sediments (e.g. Buresh and Patrick 1978; Fazzolari et al. 1990; Koike and Hattori 1978; Smith and Zimmerman 1981). Dissimilatory ammonium production is catalyzed by a variety of enteric bacteria, Bacillus species, and Clostridium species in soil (e.g. Caskey and Tiedje 1980; Prakash and Sadana 1973; Rehr and Klemme 1989; Smith and Zimmerman 1981). As with denitrification, nitrous oxide has been found to be a product of metabolism of dissimilatory ammonium producers, but the nitrous oxide appears to be produced as a side reaction in these organisms (e.g. Smith et al. 1983).
14.2 Biochemical Properties of Denitrification
During biological denitrification, nitrate is transformed to dinitrogen by a series of reductive reactions as follows:
Nitric oxide is generally enclosed in brackets in depiction of this reaction sequence since it is not usually detected as a free intermediate. There is also some controversy over whether nitric oxide is a true intermediate in the process or whether its presence represents a side reaction (e.g. Betlach and Tiedje 1981; Garber and Hollocher 1981, 1982; Hochstein and Tomlinson 1988; Hollocher et al. 1980). In either case, the significant aspects of the reaction sequence from the soil microbiologist's viewpoint are that nitrate is reduced to nitrous oxide and dinitrogen with the possible transient accumulation of nitrite plus the occasionally detected transient accumulation of nitric oxide (e.g. Cady and Bartholomew 1960; Cooper and Smith 1963). As will be indicated below, a variable ratio – depending upon the soil chemical properties – of nitrous oxide to dinitrogen is produced.
14.2.1 Carbon and Energy Sources for Denitrifiers
The electrons for nitrogen oxide reduction are provided by oxidation of a variety of fixed carbon compounds. Most commonly, glucose is added to soil samples as a carbon source in laboratory investigations of denitrification processes, but it must be remembered that a variety of organic compounds are catalyzed during this process. This array of oxidizable substrates includes many complex organic compounds, some of which are of concern from the view of soil contamination, such as acetone (Platen and Schink 1989); anthranilic acid (Braun and Gibson 1984); cresol (Bossert and Young 1986); and toluene (Evans et al. 1992). For a review of this topic, consult Evans and Fuchs (1988). Since the electrons for these oxidations are transferred through a chain of acceptors similar to the pathway leading ultimately to the reduction of molecular oxygen by aerobic heterotrophs, the growth yields of microbes oxidizing organic carbon substances while using nitrogen oxides as terminal electron acceptors are comparable with the two classes of acceptors (molecular oxygen and nitrogen oxides).
Using the traditional definition of anaerobic processes (i.e. energy‐yielding catabolic reactions obligatorily occurring in the absence of molecular oxygen), dentrifiers could be considered to be facultative anaerobes. Historically, many soil microbiologists have not considered denitrification to be a true anaerobic process in that the primary overall difference between nitrate respiration and oxygen‐based respiration is that in the former electron transport process, nitrate is merely replacing oxygen as the final electron acceptor. Denitrifiers are therefore generally thought to be aerobes. They use molecular oxygen as a terminal electron acceptor and do not ferment organic substances in the absence of molecular oxygen. They simply also possess cytochromes that can transfer electrons to nitrogen oxides. For example, in the process catalyzed by Paracoccus denitrificans terminal electron transfer to nitrate, nitrite and nitrous oxide are provided by cytochromes (e.g. Ballard and Ferguson 1988; Bolgiano et al. 1989) (Figure 14.1). If oxygen is present, different cytochromes are used for its reduction to water and dissimilatory nitrate reduction is inhibited (e.g. Hernandez and Rowe 1987).
14.2.2 Induction of Synthesis of Nitrogen Oxide Reductases
A further modification of cellular metabolism of the denitrifier growing in aerobic environs is that the requisite enzymes for denitrification are not synthesized or are produced in limited quantities. Nitrogen oxidase reductase synthesis is repressed by molecular oxygen. Once the oxygen is depleted in the microsite where the putative denitrifier resides, synthesis of these denitrifier enzymes is induced. Thus, development of a maximally functional denitrification system once anoxic conditions are imposed requires about 40 minutes to three hours. For some denitrifiers the time lapse necessary for synthesis of nitrogen oxide reductases is reduced by an induction of synthesis of the first enzyme in the pathway, nitrate reductase, as the available molecular oxygen supply is being depleted. Full induction of the pathway occurs with complete exhaustion of molecular oxygen resources. Thus, the microbe is able to respond to the absence of molecular oxygen by reducing nitrate to nitrite but it avoids synthesis of the full array of denitrification enzymes should anaerobiosis not be achieved. This differential synthesis of denitrification enzymes can result in a transient accumulation of nitrite.
Figure 14.1 Summary of the electron transport pathway leading from oxidation of fixed carbon compounds to nitrate, nitrite, or nitrous oxide as terminal electron acceptors.
14.3 Environmental Implications of Nitrous Oxide Formation
At one time, the transient accumulation of nitrous oxide in reaction vessels wherein biological denitrification was occurring was ignored or considered to be a product of the closed reaction system. It is now known that in native terrestrial environments, nitrous oxide is a common product of denitrification. For example, fluxes of nitrous oxide from soils have been shown to range from 7 to 165 kg N ha−1 yr−1 in drained histosols in South Florida (Terry et al. 1981a). Nitrous oxide evolution from mineral soil‐based systems is usually less than reported from these drained organic soils. Annual losses from irrigated soils in California range from 19.6 to 41.8 kg N ha−1 (Ryden and Lund 1980). Lesser yields have been detected from cropped soils (Benckiser et al. 1996; Goodroad et al. 1984; Mosier and Hutchinson 1981) and prairie soils (Mosier et al. 1981). Intensity and nature of fertilization used in cropping systems also may affect the quantity of nitrous oxide produced (Loro et al. 1997; McTaggart et al. 1997; Mulvaney et al. 1997). Noncropped, native ecosystems also produce more limited quantities of nitrous oxide during plant growth seasons (Goodroad and Keeney 1984).
These data show that nitrous oxide production by soil processes is highly variable. Soil properties controlling the quantity of this product by soil microbes include soil redox potential (Letey et al. 1981; Smith et al. 1983), moisture tension (Davidson and Swank 1986; Freney et al. 1979; Regina et al. 1996; Terry et al. 1981b), nitrate and oxygen concentrations (Blackmer and Bremner 1978, 1979; Firestone et al. 1979, 1980), and the time of day that the measurements are taken (Blackmer et al. 1982). Generally, the proportion of gaseous nitrogen products composed of nitrous oxide increases with increasing soil acidity, reduced soil temperature, and augmented soil nitrate levels, but the exact ratio of nitrous oxide to dinitrogen for a specific soil system varies with the combination of the soil chemical and physical properties existent therein and is therefore not reliably predicted.
Figure 14.2 Reaction sequence leading to ozone formation and the role of nitrous oxide in its destruction.
Understanding the variation of nitrous oxide in soil ecosystems has greater importance than simply allowing a better comprehension of soil nitrogen transformations. Nitrous oxide is an ozone‐depleting greenhouse gas. The reactions involved in this process are presented in Figure 14.2. It must be noted that not all of the nitrous oxide evolving from the soil surface results from activity of denitrifiers. A variety of soil heterotrophs as well as autotrophic nitrifiers produce this gas as a metabolic product or by‐product (e.g. Blackmer et al. 1980; Bleakley and Tiedje 1982; Bollag and Tung 1972; Davidson et al. 1986; Smith and Zimmerman 1981). Therefore the soil properties affecting nitrous oxide fluxes do not correlate only with those favoring biological denitrification. Complexity in interpreting the data is added by the fact that nitrous oxide is also produced by chemodenitrification as well as the various biological sources (see review by Bremner 1997 and examples of quantification of the processes in the field by Heinrich and Haselwandter 1997 and Nielsen et al. 1996).
14.4 Microbiology of Denitrification
14.4.1 Assessment of Soil Denitrifier Populations
The population density of denitrifiers in soil frequently has been estimated through use of most probable number (MPN) procedures (e.g. Focht and Joseph 1973; Volz 1977). These methods are based on detection of the conversion of nitrate or nitrite to gaseous end‐products in liquid culture media (Figure 14.3). MPN methods are inherently imprecise (see Chapter 2 for a discussion of the limitations of MPN procedures). More accurate and sensitive methods for enumeration of these organisms are now available as the characteristics of the genome encoding denitrification and the similarities in these gene sequences between the various denitrifiers present in soil samples are elucidated, thereby enabling the utilization of polymerase chain reaction (PCR) procedures for the quantification of denitrifiers in soil (e.g. Linne von Berg and Bothe 1992; Ye et al. 1993).
Figure 14.3 Outline of a most probable number procedure for estimating denitrifier populations in soil.
Source: Adapted from Focht and Joseph (1973).
Denitrifier population density data do provide an indication of the number of organisms present in a soil ecosystem that are capable of catalyzing the process. Unfortunately, in most soils, population densities of denitrifiers do not correlate with denitrification rates. Soil microbial communities commonly contain several million denitrifiers per gram dry soil. Because of the fact that the respiration of these organisms is not limited solely to nitrate reduction (they can also reduce molecular oxygen), their presence in the system only indicates a potential for denitrification should conditions favor its occurrence. The denitrifier populations occurring in any particular soil with an oxygen‐containing atmosphere may have developed through oxygen‐based respiration.
14.4.2 General Traits of Denitrifiers
Denitrifiers are a biochemically and taxonomically diverse group of bacteria. Although some denitrifiers are chemoautotrophs (e.g. using hydrogen or reduced sulfur compounds as energy sources) and others are photoautotrophs, most of these organisms generally derive their energy from oxidation of fixed carbon substrates, including single carbon compounds. The primary substrate and end‐product of the process are nitrate and dinitrogen, but some denitrifiers can only reduce nitrite (nitrite‐dependent denitrifiers) whereas others lack nitrous oxide reductase, thereby producing nitrous oxide as the terminal product. The variety of metabolic types of bacteria capable of denitrifying defies any effort to group these organisms into one or a few bacterial genera.
Further difficulty with evaluation of the occurrence of the “denitrifier trait” in classical literature results from the fact that many organisms classified historically as denitrifiers are not really denitrifiers. Care must be exercised in classing organisms as denitrifiers based strictly on the conversion of nitrate or nitrite to nitrous oxide or dinitrogen independent of the extent of the reaction. Many bacterial strains are capable of reducing nitrate and of producing limited quantities of dinitrogen or nitrous oxide. Although some of these organisms have been termed denitrifiers in the past, these marginal nitrate reducers are not true denitrifiers. To be a true denitrifier, a microbe must meet the following criteria.
· At least 80% of the nitrate or nitrite reduced by the bacterium must be converted to dinitrogen and nitrous oxide.
· There must be an increased growth yield due to the reduction of nitrate, nitrite, or nitrous oxide. This trait is the primary requirement for classing an organism as a denitrifier in that it shows that nitrogen oxide reduction is a dissimilatory process.
· The conversion of nitrate to nitrous oxide and dinitrogen must occur at a high rate. That is, the process must be central to cellular intermediary metabolism, not just a side reaction providing a minor pathway for electron transport.
· The presence of cytochrome cd or dissimilatory nitrite reductase should be demonstrable in the microbial cells.
14.4.3 Generic Identity of Denitrifiers
Many of the denitrifiers isolated from soils and waters are members of the genera Pseudomonas or Alcaligenes, but a large number of other bacterial genera contain strains of denitrifiers (e.g. Gamble et al. 1977). As with diazotrophs, denitrification is not a sufficiently definitive trait to allow separation of all denitrifiers into one or a limited number of genera. Not only are the individual bacterial strains of denitrifiers metabolically distinct, but the properties of the enzymes associated with denitrification process itself are highly variable. For example, in a comparison of denitrification by Pseudomonas stutzeri, Ps. aeruginosa, and Ps. denitrificans it was found that the rates of anoxic growth of the organisms varied 1.5‐fold, gas production varied over eightfold, and cell yield differed by threefold (Carlson and Ingraham 1983). The metabolic and taxonomic diversity of these organisms is exemplified by the following partial listing of some of the bacterial genera containing denitrifiers (for a more complete listing of denitrifiers, see Knowles 1981, 1982).
· The genus Achromobacter is cited as containing a variety of denitrifiers, including methane oxidizers and nitrite‐dependent denitrifiers. (These organisms have now been grouped among the species of the genus Alcaligenes.)
· Alcaligenes eutrophus, an organism originally classified as Hydrogenomonas eutrophus, is capable of autotrophic growth using hydrogen as an energy source, carbon dioxide for carbon, and nitrate as a terminal electron acceptor (e.g. Pfitzner and Schlegel 1973). This bacterial species is a facultative autotroph since it also uses organic carbon as an energy source. Other common denitrifiers in the genus Alcaligenes include A. denitrificans and A. odorans. Another facultative autotroph which oxidizes hydrogen for energy is Ps. denitrificans.
· Several nitrogen‐fixing bacteria are also capable of denitrifying. These diazotrophs include some strains of Azospirillum brasilense (e.g. Chauret et al. 1992; Neyra et al. 1977) and some Rhizobium species (e.g. Chan et al. 1989; Coyne and Focht 1987; Daniel et al. 1980; van Berkum and Keyser 1985; Zoblotowicz et al. 1978)
· Some denitrifiers are also thermophilic (e.g. Hollocher and Kristjánsson 1992).
· Some strains of Chromobacter denitrify using nitrate or nitrite as terminal electron acceptors, reducing them to nitrous oxide and dinitrogen, whereas Chromobacter violaceum can only reduce nitrogen oxides to nitrous oxide (Bazylinski et al. 1986).
· Halobacterium marismortui is a halophilic denitrifier that was isolated from the Dead Sea.
· Hyphomicrobium strains denitrify with methanol as the primary energy source. The practical significance of understanding the environmental limitations to activity of these strains relates to the fact that methanol has been used as an additive to sewage effluent to provide the energy source necessary for maximization of denitrification rates for removal of nitrate from the effluent.
· The photosynthetic bacterium Rhodopseudomonas sphaeroides forma sp. denitrificans is both a denitrifier and a diazotroph (Dunstan et al. 1982).
· Thiobacillus denitrificans is an example of a sulfur oxidizing chemoautotroph which grows in the absence of oxygen with nitrate as a terminal electron acceptor. The energy sources for this organism include sulfide, elemental sulfur, and thiosulfate. T. denitrificans can also use molecular oxygen as a terminal electron acceptor.
14.5 Quantification of Nitrogen Losses from an Ecosystem via Denitrification
Historically, the most commonly used techniques for assessing nitrogen fixation have involved (i) determination of nitrogen balances in the ecosystem and (ii) quantification of distribution and movement of 15N‐labeled fixed nitrogen through soil reservoirs. More commonly used laboratory‐based assays of denitrification are (i) assessment of disappearance of nitrogen oxides from soil slurries and (ii) use of acetylene inhibition of nitrous oxide reductase. The principles, advantages, and deficiencies of each of these techniques will be assessed herein.
14.5.1 Nitrogen Balance Studies
As was discussed previously in relation to quantification of nitrogen fixation in field sites, assessment of changes in nitrogen distribution among the various soil reservoirs as a means of quantifying process rates appears to be conceptually simple and logically attractive. For application of nitrogen balance methods to quantification of denitrification, the amount of nitrogen denitrified is concluded to be the unaccountable or missing nitrogen. The quantity of nitrogen denitrified is that preexisting in the system plus any inputs (e.g. nitrogen fixation, fertilizer inputs) minus known losses (e.g. nitrogen in crops, nitrogen leached, ammonium volatilized). The procedure is also called the difference method because all nitrogen occurring in the system is subtracted from the sum of that previously occurring therein plus any amendments.
Use of the nitrogen balance method for estimation of denitrification processes in soils is predicated upon three assumptions.
· All of the organic and fixed nitrogen compounds contained in soil can be accurately quantified.
· The assay procedures available for quantifying soil nitrogen compounds are sufficiently sensitive to allow for detection of small changes in the various soil nitrogen compounds. Furthermore, it is assumed that the soil, water, and air samples analyzed are representative of the ecosystem and that no changes in the nitrogen levels existing in these samples at the site occurred between collection and analysis.
· The sole means of loss of gaseous nitrogen from soil is denitrification.
Unfortunately, none of these assumptions is valid. Reasonably sensitive assay procedures are available to assay soil organic and inorganic nitrogen pools. Their application to soil nitrogen cycle analysis is limited by the fact that changes in soil nitrogen reserves are commonly less than the precision of the assay method. Further problems with the direct quantification of nitrogen species in a particular soil site result from the fact that soil is an open system. That means water‐soluble substances such as nitrate, nitrite, and simple organic nitrogenous compounds move freely into and out of the portion of the ecosystem of interest. Inability to accurately account for all fixed nitrogen inputs and losses from the soil site results in appreciable statistical variability in nitrogen balance data. This variation also results in invalidation of the second assumption.
Probably the biggest difficulty with application of nitrogen balance studies to assessment of denitrification is the third assumption (all nitrogen volatilization results from denitrification). There are at least three major routes of nitrogen volatilization from soil. Nitrogen may be lost through:
· nonbiological losses of ammonia (ammonia volatilization)
· chemical decomposition of nitrite
· denitrification as dinitrogen and nitrous oxide.
Ammonia volatilization can be a major route for loss of fixed nitrogen from soil. Up to 25% of the ammonia applied to soil as fertilizers or formed microbiologically may be lost as ammonia. Fortunately, for most soil systems, this process of nitrogen volatilization is a minor concern since ammonia losses are insignificant below pH 7.0. (Recall that at acidic pHs the equilibrium between ammonium and ammonia is shifted so that ammonium predominates.) Ammonia volatilization is particularly apparent in the use of farmyard manure and at times in association with fertilization with urea. Major environmental concern with this process is linked more to the migration of soil fixed nitrogen to nearby bodies of water than with its interference with quantification of denitrification.
Nitrite decomposition (chemodenitrification) is rarely a major route of exodus of fixed nitrogen from soils. The most commonly documented situation for accumulation of meaningful nitrite concentrations is linked to alkaline soil conditions where nitrite accumulates due to ammonia toxicity to Nitrobacter spp. This phenomenon occurs in soils fertilized with anhydrous ammonia. In this process, nitrite reacts chemically with soil organic matter to produce dinitrogen and nitrous oxide plus some nitric oxide. In the absence of molecular oxygen, the nitric oxide may be evolved, but it is generally oxidized to nitrogen dioxide as soon as it is exposed to molecular oxygen. Although this reaction can occur in soil, nitrite will more commonly be biologically denitrified in soil. Thus, the relative importance of the three pathways of nitrogen volatilization from soils may be ranked, with denitrification being the major means of nitrogen loss, ammonia volatilization being ranked a distant second, and nitrite decomposition occurring only as a minor contributor.
The uncertainties associated with the use of nitrogen balance procedures for estimation of soil nitrogen losses due to denitrification were exemplified by data from a study in which this procedure was compared to results from a parallel quantification of nitrogen dynamics using 15N to trace changes in soil nitrogen reservoirs (Rolston et al. 1979). In the more moist soils studied, actual denitrification losses were generally 0–30 kg N ha−1 less than amounts calculated from the nitrogen balance procedures. These differences ranged from 12 to 65 kg N ha−1 for drier soils. The authors concluded that with their soils, the uncertainties associated with the nitrogen balance procedure resulted primarily from large sampling variability in nitrogen leaching results (although variability of soil residual nitrate and organic nitrogen was also large). These results further emphasize the conclusion that even in soils where nitrogen volatilization due to chemical processes is minimal, the accuracy of the nitrogen balance procedure is limited due to difficulties in precisely accounting for all soil nitrogen forms present and leaving the soil system.
14.5.2 Use of Nitrogen Isotopes to Trace Soil Nitrogen Transformations
Although the short‐lived radioactive nitrogen isotope (13N) has been used to quantify denitrification processes (Firestone et al. 1979; Hollocher et al. 1980; Spier et al. 1995a, b; Tiedje et al. 1979), due to the short half‐life of the radioactive isotope, the heavy isotope of nitrogen (15N) is more amenable to long‐term quantification of denitrification reactions in soils. For these studies, 15N‐labeled fixed nitrogen sources (organic or inorganic) are amended to field soils or incubated soil samples (greenhouse or laboratory) and the 15N‐labeled dinitrogen or nitrous oxide quantified.
As with the evaluation of biological nitrogen fixation, implementation of this procedure is predicated on a series of assumptions, the most significant being (i) the denitrifier populations do not discriminate between nitrogen isotopes and (ii) the soil‐amended 15N‐labeled fixed nitrogen is homogeneously mixed with native soil nitrogen resources and is thus equally available to the denitrifier population. As was noted with biological denitrification, satisfaction of these assumptions is questionable.
Discrimination of nitrogen isotopes by denitrifiers has been demonstrated (e.g. Blackmer and Bremner 1977). Clear microbial capacity to discriminate between 14N‐ and 15N‐labeled nitrate was observed by these workers. Significant nitrogen isotope effects were detected in both reduction of nitrate to nitrite and nitrite to gaseous end‐products.
Furthermore, because of the heterogeneity of the soil system, uniform mixing of an external nitrogen supply with indigenously formed fixed nitrogen is difficult at best. Vanden Heuvel et al. (1988) found that the magnitude of this error increases with the range of 15N enrichment between native soil N isotope distributions and amended nitrogen (for a constant mean 15N enrichment) and it varies depending on the number of nitrate pools existent in the system. The error was not affected by overall nitrogen gas evolution rate or the 15N enrichment of the amended fertilizer. The theoretical basis for this potential for overestimation and underestimation of nitrous oxide and dinitrogen evolution was documented through mathematical modeling by Boast et al. (1988).
Even though these cited studies demonstrate significant difficulties with utilization of 15N to quantify losses of fixed nitrogen from soil ecosystems due to denitrification, its application to field, laboratory, and greenhouse studies is appropriate when the limitations are considered. The use of 15N is appropriate for field experiments (especially for long‐term studies as exemplified by Focht and Stolzy (1978)) when the limitations are considered during data interpretations. 15N labeling of nitrogen species in greenhouse and laboratory incubation studies has also been particularly useful. In the latter situations, extensive disruption of soil structure during sample preparation is common and the conditions for denitrification are more controlled.
14.5.3 Soil Nitrogen Oxide Transformations
Since nitrate, nitrite, and nitrous oxide are readily extracted from soil and quantified, it is reasonable to consider that the rate of change in concentrations (native or amended) of these chemicals can serve as an indication of denitrification rates in soils (e.g. Bremner and Shaw 1958a) in laboratory‐incubated soil samples. This procedure is based on the assumption that all of the nitrate or nitrite lost from the system is denitrified. Except in systems where extensive dissimilatory ammonium production is occurring, alternative fates of the nitrogen oxide (generally nitrate) tend not to interfere with this analysis procedure. Clearly, some nitrate could be incorporated into microbial biomass if the assay is conducted over a timeframe sufficiently long for microbial growth to occur. Thus, in this situation, denitrification rates would be underestimated by the amount of nitrate assimilated by the microbial biomass. Since the assays are conducted under laboratory conditions, plant uptake and leaching are not important because plants are excluded from the system and the reaction vessel is a closed system.
Greater problems with the use of changes in nitrogen oxide levels as an assessment of denitrification kinetics in soil are derived from the fact that in most soil systems, nitrate concentrations are a few μg g−1 soil at best. Generally, nitrite is not detected in native soil samples. Thus, soils are generally amended with exogenous nitrate or nitrite for a meaningful estimate of the capacity of the soil microbial community to denitrify nitrogen oxides to be attained. To assure that the totality of the denitrification enzymes is measured (that is, enzyme and not substrate is the limiting factor), enzyme saturating concentrations of nitrate or nitrous oxide are generally added to soil with this technique. Therefore, since these nitrogen oxide concentrations are artificially elevated above those normally occurring in the field, denitrification rates assessed using this procedure must be understood to be potential rates rather than actual field values. The potential values would only be equivalent to field denitrification rates in the unusual situations when field nitrate concentrations are sufficient to saturate denitrification enzymes present.
14.5.4 Acetylene Block Method for Assessing Denitrification Processes in Soil
Incubation of soil samples or amendment of the soil atmosphere in situ with 10−2 atm of acetylene inhibits the reduction of nitrous oxide to dinitrogen (Yoshinari and Knowles 1976; Yoshinari et al. 1977). Since commonly available gas chromatography procedures may be used to quantify nitrous oxide, this provides a sensitive method for measurement of denitrification enzymes. (See Figure 14.4 for an outline of the procedure.) The validity of this technique for determination of denitrification kinetics in soil slurries (e.g. Smith et al. 1978) and in field situations (e.g. soil cores [Parkin et al. 1985a] and direct field measurements [e.g. Aulakh et al. 1991; Mosier et al. 1986]) has been confirmed through the comparison of data derived from use of acetylene block methods with those from 15N and 13N enrichments. Acetylene reduction inhibition of nitrous oxide reduction is a quick and reliable procedure for estimation of in situ levels of denitrification enzymes in soil slurries, soil cores, and the field. The initially constant nitrous oxide evolution rate in soil slurries corresponds to preexisting denitrifier enzyme in the soil sample (Smith and Tiedje 1979a,b).
Figure 14.4 Outline of a general procedure using the acetylene block method for quantifying denitrification potential in soil samples incubated in the laboratory. Samples receiving nitrate amendment provide estimates of maximum denitrification potential of the sample. Incubation times may vary from one hour to several days, depending upon the experimental objectives.
A problem of providing an inhibitory level of acetylene throughout the soil sample is most frequently encountered in field sites and, to some degree, soil column experiments. In the field situation and with soil columns, care must be taken to assure complete admixture of the acetylene gas with the soil atmosphere. This limitation to the assay procedure is usually not important when using soil slurries since the suspensions are agitated at sufficient rates to guarantee distribution of the inhibitory gas through out the mixture.
Acetylene is toxic to ammonium oxidizers (e.g. Berg et al. 1982). Thus, in long‐term field or laboratory studies where indigenous nitrification is relied upon for the nitrate source for denitrifiers, denitrification would be reduced or precluded. Implementation of this procedure in the field or with long‐term incubated soil columns requires either duplication of sampling sites so that assays would not be repeated at the same specific site receiving acetylene or sufficient time must be allowed between assays for recovery of the nitrifier population to occur.
Additional difficulties may be encountered in the field should the acetylene itself stimulate carbon mineralization, thereby augmenting nitrate reduction (Haider et al. 1983). Nitrate amendment to soil samples may stimulate both aerobic and anaerobic acetylene oxidation. Based on this observation, Haider et al. (1983) conclude that repeated application of acetylene to the same field plot should be avoided.
It must be noted that these complications are not universally associated with field implementation of the acetylene inhibition procedure. Ryden and Dawson (1982) found no difficulty with its repeated use on grassland soils. In conclusion, these data further underscore the necessity of conducting appropriate controls to assure that the impact of acetylene on the soil system is direct alteration of nitrous oxide evolution through inhibition of nitrous oxide reductase and not indirectly through alteration of the activity of microbial populations dissociated from nitrate reduction.
Two additional factors to be considered in regard to the use of acetylene to estimate denitrification potential in soil is the capacity to estimate nitrous oxide production from both autotrophic nitrifiers and denitrifiers in the same soil sample and the existence of potential difficulties due to enhanced nitric oxide decomposition in soil receiving acetylene. Since nitrifiers and denitrifiers are inhibited by different concentrations of acetylene, variations in both concentration of acetylene and duration of exposure (Kester et al. 1996) or concentration (Inubushi et al. 1996; Webster and Hopkins 1996) have proven valuable in estimating the quantity of nitrous oxide produced by each source in soils and sediments. (Alternatively, Stevens et al. 1997 used differential 15N‐labeling of nitrate and ammonium pools to distinguish the two nitrous oxide sources.)
An additional concern in situations where nitric oxide is a by‐product of denitrification is the impact of acetylene on the decomposition of the nitric oxide. Increasing acetylene levels in soil have been shown to increase the decomposition of nitric oxide by as much as 5–557‐fold (Bollmann and Conrad 1997a). The consequence of the augmented nitric oxide decomposition is an underestimation of the associated denitrification rate (Bollmann and Conrad 1997b).
14.6 Environmental Factors Controlling Denitrification Rates
The primary factors that have been shown to control denitrification processes in the field are (i) the nature and amount of organic matter available as energy sources to the denitrifiers, (ii) the soil nitrate concentration, (iii) the aeration/moisture status of the soil, (iv) soil pH, and (v) soil temperature. Each soil property and examples of their effect in native soil systems will be evaluated herein.
14.6.1 Nature and Amount of Organic Matter
Most denitrifiers in soil are heterotrophs. Thus, the primary energy source for the process is plant biomass in most soil ecosystems. Although colloidal organic matter is a primary soil component, amendment of soil with metabolizable carbon generally does stimulate denitrification (e.g. de Catanzaro and Beauchamp 1985; Stanford et al. 1975b). For such experiments, saturating concentrations of nitrate are usually amended to soil slurries with the carbon and energy source (frequently glucose) varied. The increase in denitrification rate due to the augmented energy source generally resembles Michaelis kinetics (e.g. Bowman and Focht 1974; Reddy et al. 1982). Bowman and Focht (1974) found a Michaelis constant of 500 μg glucose mL−1 when nitrate reduction was assessed in a Coachella fine sand. Further evidence for carbon limitations of denitrification rates is derived from the observation that denitrification rates generally correlate with total organic carbon and available carbon levels in soil. Available carbon has commonly been estimated by measurement of biological oxygen demand of soil water extracts (e.g. Beauchamp et al. 1980; Burford and Bremner 1975; Katz et al. 1985; Stanford et al. 1975c). There are rare situations in field sites where carbon does not limit denitrification. This exception is usually associated with mineral soils receiving high organic matter inputs (e.g. sludge‐amended soils).
The generally observed correlation of denitrification rates with available organic matter leads to the prediction that treatment of soil to increase availability of native soil organic matter should result in an increase in the denitrification rate. Therefore, soil manipulations that alter the accessibility of indigenous organic matter to microbes can also change the indigenous denitrification rate measured. Freezing (e.g. McGarity 1962) and air drying (e.g. Patten et al. 1980) of soil samples can augment the denitrification rate detected with the samples. Thus, even with the most careful handling of field soil samples, the possibility that the laboratory‐derived results are an overestimation of field denitrification rates must be considered.
An interesting possibility based on the observed proportionality of denitrification rates to available fixed carbon substrates is that the denitrification rate may be elevated in the rhizosphere compared to nonrhizosphere soil. Root exudates can be predicted to provide the energy source for the denitrifiers and anoxic microsites would be anticipated to occur in the vicinity of the roots where denitrifiers can function. Addition of macerated roots to soil samples as well as invasion of soil by root tissue has been shown to stimulate denitrification (e.g. Garcia 1975; Prade and Trolldenier 1988). Others (Haider et al. 1985, 1987; Smith and Tiedje 1979b) did not observe a rhizosphere effect on denitrification. Since the denitrifiers and the plants could be competing for the same nitrate pool, it could be hypothesized that the reason for the disparate observations is that different nitrate concentrations were used in the studies but Tiedje et al. (1979) found that variation of soil nitrate concentrations did not alter the denitrification rate in their rhizosphere samples. No rhizosphere effect on denitrification rates was detected with high nitrate concentrations. A reduction of denitrification occurred in the presence of low nitrate. Since the reduced nitrate concentrations used in their studies better reflected the generally encountered field situations, it could be predicted that denitrification rates in the rhizosphere should be reduced over those of nonplant‐impacted soils.
These data exemplify the complexity of the interactions of physical, chemical, and biological factors in controlling denitrification in the rhizosphere. Although there is a clear impact of fixed carbon concentration on activity of denitrifiers, their function is also delimited by nitrate concentrations (for which they would be competing with plants as well as other bacterial populations), and anoxic microsites.
14.6.2 Nitrate Concentration
Since denitrification is an enzymatically catalyzed process, the reaction rate of the process is anticipated to follow Michaelis–Menten kinetics. That is, the rate of nitrate reduction to dinitrogen and nitrous oxide should increase until a saturating concentration is reached. Once enzyme saturating concentrations of the substrate have been reached, two alternatives are generally observed. There can be no effect of the augmented substrate levels or the substances may become inhibitory. The latter situation is generally detected in study of denitrification. That is, high nitrate concentrations tend to be inhibitory to nitrogen oxide reductases.
Nitrate reduction in soil is generally found to be either a zero‐order or first‐order process. For example, McGarity (1961) found denitrification rate to be independent of nitrate concentration (60–472 μg g−1) in some South Australian soils. Bowman and Focht (1974) observed first‐order kinetics with a Coachella fine sand. The Michaelis constant for the latter study was 170 μg nitrate mL−1. Murray et al. (1989) measured Michaelis constants for mixed denitrifier populations from agricultural soils of 1.8–13.7 μM nitrate. Combinations of kinetic relationships may also be observed. Reddy et al. (1978) measured both first‐order and zero‐order reaction kinetics with flooded soils. In their study, nitrate disappearance from the waters above the soil was zero order whereas within the soil matrix, nitrate consumption followed first‐order kinetics. In this situation, a nonbiological parameter entered into the reaction kinetics – diffusion of nitrate from the surface waters into the sediment soils. Diffusion limited the quantities of nitrate available to the soil microbes so the available nitrate was reduced by the microbial community as quickly as it entered into their microsite. Myrold and Tiedje (1985) suggest limitations to diffusion of nitrate to denitrifiers in large soil aggregates.
At high concentrations, nitrate inhibits nitrous oxide reductase. This inhibition varies with soil pH and history of the soil. Greater effect of high nitrate is detected at low pH, whereas extended flooding of the soil relieves the inhibitory effect of nitrate on the reductase.
14.6.3 Aeration/Moisture
As was noted above when the biochemistry of denitrification was analyzed, oxygen represses the synthesis of the reductases associated with denitrification. As soil oxygen tensions are reduced, reduction of nitrate to nitrous oxide and dinitrogen increases (e.g. Allison et al. 1960; Cady and Bartholomew 1961). Because of the slow rate of diffusion of molecular oxygen in water compared to air, this effect of aeration on denitrification rates in soils is highly related to soil moisture. In general, denitrification is not detected below 60% water‐holding capacity. Above this value, denitrification rate generally correlates with soil moisture.
The field implication of this observation is that it can be anticipated that the majority of nitrate is denitrified in soil under flooded conditions. Generally, low denitrification rates are observed continuously in a drained soil system, but maximal activity is associated with increases in moisture due to rainfall (Sexstone et al. 1985a). In the field as soil moisture increases, loss of nitrate through denitrifier activity increases (e.g. Kroeckel and Stolp 1988; Pilot and Patrick 1972). This sensitivity of denitrifiers to increasing soil moisture explains a significant amount of the augmented loss of nitrate from no‐till soils due to denitrifier activity compared to that occurring in conventionally tilled agricultural systems (Aulakh et al. 1984a, b; Rice and Smith 1982). This increase in denitrification results from augmentation of the number of anaerobic microsites in the soil. That is, the low amount of denitrification occurring in drained soils is occurring in anaerobic microsites and is not the result of activity of molecular oxygen‐resistant denitrifiers (Hojberg et al. 1994; Sexstone et al. 1985b). Anaerobiosis is not the sole factor controlling the denitrification rate in the aggregate in that this rate does not correlate with the size of the anaerobic zone. Other contributing parameters are nitrate diffusion rate into the anaerobic microsites and the rate of molecular oxygen consumption by the aerobic organisms growing on the surface of the aggregate.
14.6.4 pH
Although denitrification occurs within the pH range of approximately 3.9–9.0, maximum nitrogen oxide reduction occurs from pH 7.0 to 8.0. The rate decreases as the pH is lowered (e.g. Bremner and Shaw 1958b; Focht 1974; Waring and Gilliam 1983). Acidophilic or acid‐tolerant denitrifier populations are apparently selected in soils with histories of low pH. Parkin et al. (1985b) noted two distinctly different pH optima with soils of pH 3.9 and 6.3 that approximated the native pH values of the soils. It should be noted that the acidic soil used in the latter study had a 20‐year history of low pH which would have provided more than adequate time for development of an acid‐resistant soil microbial population.
Because of the occurrence of chemical denitrification in extremely acidic soils (e.g. Bollag et al. 1973), assessment of denitrification in such soils is difficult. Fortunately, the major products of the chemical and biological processes differ. Nitric oxide and nitrogen dioxide predominate from chemical reduction of nitrite whereas nitrous oxide and dinitrogen are biologically produced (Bollag et al. 1973). Unfortunately, dinitrogen and nitrous oxide are also reported among the products of chemodenitrification (e.g. Bulla et al. 1970; Reuss and Smith 1965). When the two processes are separated significant biological reduction of nitrous oxides can be shown to occur. Muller et al. (1980) found biological denitrification rates varied from 0.12 to 53.8 μg d−1 in a variety of low pH (pH ≥ 3.6) spodosols and peats collected from southern Finland. Complete inhibition of the reduction of nitrous oxide to dinitrogen was found in 99.3% of these soil samples. Similarly, Gilliam and Gambrell (1978) found that acidity (pH values as low as 4.5) was not a serious limitation to denitrification in Atlantic coastal plain soils (U.S.A). As with the previous study, nitrous oxide was a major product of denitrification in these soils (Weier and Gilliam 1986). Similarly, in a study of a silt loam soil with pH varying from 4.6 to 6.9, nitrous oxide made up 83% of the gaseous nitrogen products of denitrification at pH 4.6 and 5.4. At pH 6.9, dinitrogen was the predominant product.
Frequently, several soil physical or chemical properties can interact to retard denitrification under acidic conditions. For example, Dubey and Fox (1974) found that low pH combined with low soil organic matter precluded denitrification in humid tropical soils of Puerto Rico. Similarly, George and Antoine (1982) noted that the pH optimum varied with temperature in their studies of a salt marsh soil. In their study, nitrous oxide was produced only at low pH and nitrate concentrations.
14.6.5 Temperature
The minimum temperature for denitrification is generally associated with the occurrence of free water. Smid and Beauchamp (1976) extrapolated their data from study of denitrification in the A horizon of a Huron clay loam to suggest denitrification at or near 0 °C. Incubation of the same soil under anaerobic conditions resulted in a complete inhibition of denitrification at 5 °C with a gradual increase in the rate to 30 °C (Bailey and Beauchamp 1973). Jacobson and Alexander in a study of two soils maintained in an anaerobic atmosphere (Jacobson and Alexander 1980) found no nitrate‐reducing activity at 1 °C. Nitrate was slowly reduced at 7 °C. As with the previous studies, the rate increased with increasing soil temperature. These data suggest a minimum temperature for denitrification approaching the freezing point for water with an apparent impact of the presence of molecular oxygen on the limiting temperature. The maximum temperature is approximately that value limiting biological activity in soil in general, 75 °C. The optimum temperature (60–70 °C) is well above the normally occurring soil temperature range. In the range of increasing temperature, the nitrate reduction rate increases at a rate approximating that of general soil biological processes. That is, a Q10 of approximately 2 describes the relationship between increasing temperature and reaction rate (e.g. Stanford et al. 1975a).
These data indicate that significant denitrification can be anticipated to occur in any soil ecosystem at temperatures where general biological activity occurs. The Q10 value of about 2 indicates that the process is reasonably sensitive to slight fluctuations in soil temperature. Thus, not only should a seasonal variation in this process be anticipated, but also a diurnal impact. That is, losses of fixed nitrogen as nitrous oxide and dinitrogen will necessarily vary with daily heating and cooling of the upper portions of the soil profile.
14.6.6 Interaction of Limitations to Denitrification in Soil Systems
A strict consideration of each of the soil properties listed above and their impact on denitrification processes provides a limited view of the variation of nitrate reduction in field sites. In reality, each soil property is sufficiently variable across the soil landscape to provide significant landscape‐scale variation in denitrification rates. Landscape analysis should be incorporated into models of soil fixed nitrogen losses through denitrification since topography has a major impact on in situ denitrification rates (Ball et al. 1997; Pennock et al. 1992). In the Pennock et al. (1992) study, when the site was considered as a whole, variables that were most influential on nitrogen losses were volumetric water content and soil redox potential. In the level portions of the site, volumetric water content was most highly correlated with denitrification whereas on the shoulder and foot slopes, respiration and bulk density were the most influential. Parkin (1987) found that “hot spots” of denitrification activity were associated with pockets of particulate organic matter. Parsons et al. (1991) found coefficients of variation between 74% and 268% for spatial variability of nitrogen gas loss from agricultural soils of central Kentucky (USA). Increases in nitrate reduction activity related to increases in soil moisture and soil respiration. These data support the conclusion that for a complete understanding of denitrification kinetics in a soil system, a study of the heterogeneity of the system, particularly with regard to soil moisture, organic matter, and most likely to some degree nitrate distribution, must be developed.
14.7 Conclusions
An understanding of the nuances of the impact of soil properties on biological denitrification and management of this process to regulate soil nitrate loadings is essential for optimization of agricultural production as well as for minimization of the potential for soil and water pollution. This terminal stage in the nitrogen cycle (return of fixed nitrogen to atmospheric reservoirs) is catalyzed by a diverse group of soil bacteria that are capable of using light, mineral (inorganic), or organic compounds as energy sources. The only readily discernible commonalties among denitrifiers are (i) they are solely bacteria, (ii) they are capable of reducing nitrogen oxides in the absence of free oxygen, and (iii) these nitrogen oxides serve as terminal electron acceptors for bacterial respiration. Thus, the primary determinants of denitrification rates in the field are the availability of energy sources, nitrate, and the absence of molecular oxygen from the microenvironment of the bacteria. The diversity of the organisms involved leads to the prediction that if life is possible in the system, a denitrifier should exist which is capable of functioning therein. This adaptability and ubiquity of denitrifiers supports the conclusion that denitrification is a process amenable to management for control of soil nitrate loadings.
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