Common section

9

Introduction to the Biogeochemical Cycles

Consider a handful of soil. Most soils (i.e. mineral soils) generally have the look and feel of a complex mixture of mineral particles; that is, they are composed of a reasonably simple mixture of sand, silt, and clay. This observation could easily lead to the conclusion that soil is simply a mixture of mineral particulates. Perhaps we might also conclude that variability in soil function could be derived, at least in part, from differences in chemical composition and particulate size. A bit more contemplation could lead to the conclusion that the role of soil in the ecosystem only extends to the properties associated with providing physical support for plants and directing water and air to plant roots. Over the millennia, we have come to appreciate the fact that the easily seen soil components are closely associated with myriad organic compounds and living organisms ranging from nano‐sized bacteria and viruses to easily recognized higher animal and plant life. Complex biogeochemical cycles of life requiring minerals and mineralization of decaying biomass provide the energy and nutrients required for growth of the soil biological community. Thus, this chapter has the objectives of examining the relationships held in common by organisms producing the nutrients and the processes used in their study.

As our examination of soil properties extends beyond our initial observations of the complexity of the physical structure, we must ask the simple question “Why is such complexity in structure necessary?” The answer to this query relies on considerations beyond examination of individual soil minerals. The role of soil in relationship to the function (existence) of the total ecosystem must be considered. A good starting point could be the contributions of soil to plant community growth and development. Clearly, soil must provide a physical support for plants as well as containing the basic elements required for plant growth. Some initial conclusions regarding soil structure and variability include the following.

· The nutrient needs of the plants change throughout the growing season.

· Temperature in the soil change daily.

· Soil moisture is highly variable.

· Growth of the plant changes soil chemical and physical properties.

More succinctly stated, variability in soil chemical, biological, and physical properties necessitates changes in soil properties required to maintain a vibrant, productive plant community just as the aboveground plant community must evolve to sustain its development and resilience.

Therefore, it must be concluded that the complexity of soil components and their interactions supports (i) the resistance of the ecosystem to external perturbations, (ii) the productivity of the currently existing living community, and (iii) the sustainability of the ecosystem. That is, the answer to our simple question regarding the necessity of such a complex mixture of soil particles and life could be reduced to the simple statement “no complex, evolving soil system – no terrestrial life.” This conclusion is further supported by a short list of soil services involving soil microbes, examples of the function of soil organic matter (humus), and the roles of soil aggregates as an introduction to the importance of the interaction of soil physical, chemical, and biological processes.

An important yet perhaps not obvious conclusion revealed by this introduction is that a primary driving force for the cycling of nutrients in soil is the capacity of the soil biological community to mineralize plant organic matter for the energy and nutrients required for replication. Further analysis will show that transformations supportive of the biogeochemical cycles involve external organic matter sources (plant biomass) as well as that produced through chemical and biological reactions within the soil itself.

Thus, basic questions arising at this point include “What is soil organic matter (humus) and how does it differ from the plant biomass, the primary source of nutrients and energy supporting growth of the soil biological community?” The major groupings of organic matter makingg up humus include (i) partially decomposed plant debris, (ii) microbial biomass, (iii) identifiable classes of biochemicals (e.g. cellulose, lignins, etc.) that have been separated from the cells synthesizing them, and (iv) humic substances. The complexity of organic chemicals derived from plant and animal biomass is reasonably familiar; humic substances may be somewhat less familiar. This rather cursory examination of common soil organic components leads to the conclusion that they are derived from transformations of native plant and animal biomass, humic substances formed in situ, human waste materials, and anthropogenic chemicals.

At this point of our analysis, we can conclude that the diversity of organic chemicals and their fate in the soil system presents a potentially overwhelming task for those charged with examining nutrient cycling in soils. Thus the primary objective of this chapter is to provide a basic understanding of biogeochemical cycles through evaluating the general properties of the cycles and examining the interrelationships of these processes with the goal of developing an appreciation of the importance of these processes to overall ecosystem function. Specific biogeochemical cycles are examined in greater detail in subsequent chapters.

9.1 Introduction to Conceptual and Mathematical Models of Biogeochemical Cycles

A primary maxim for this analysis of biogeochemical cycles is that the energy and nutrients used by soil microbial biomass are derived from microbial recovery of the nutrients and energy contained in aboveground plant biomass. That is, one could state that since the majority of the soil microbial biomass results from mineralization of photosynthetically produced plant biomass, the soil biological community is ultimately solar powered, with the primary pathway for energy inputs to support belowground life being solar‐powered plant biomass synthesis to soil primary producers (e.g. bacteria and fungi). Other natural sources of microbial nutrients and energy include animal biomass as well as nutrients contained in rainwater and other surface water, although to a much lower degree than the plant biomass. In systems with high levels of human intrusion, products of societal development and function also play a meaningful role in this nutrient cycling and energy transformation.

A more complete consideration of the organic substances supporting soil systems therefore requires an understanding of transformations of the biochemical components of living biomass and anthropogenically synthesized organics, as well as the more biodegradation‐resistant substances such as humic substances as well as lignin. Elucidation of the processes associated with the building blocks of microbial biomass requires study of the chemical nature, physical location within the soil system, and its degradation susceptibility. For plant biomass, some components are easily converted into microbial biomass and energy (e.g. carbohydrates and proteins), whereas others are not (e.g. lignin). Furthermore, meaningful quantities of the organic compounds entering soil are transformed biologically as well as chemically into complex organic materials (e.g. humic acid, fulvic acids, and humin) that are not easily metabolized by the soil microbial community. Understanding the nature of the transformation of soil and plant organic matter as well as a wide variety of human‐synthesized organic components amended to soil requires a basic understanding of the variable physical states of these materials within soil and the extremely variable distributions of organic substances within the soil matrix. For example, as noted in earlier chapters, organic matter may be occluded to the microbial community by its incorporation into soil aggregates, binding to clay, or being contained within resting microbial cells.

9.1.1 Development and Utility of Conceptual Models

Elucidation of the diversity of organic substances and their interactions with the soil physical structure is essential for developing appropriate soil management plans, determining sustainability of the soil system, and evaluating the impact of transformation of organic matter on external processes, such as mediating climate change. Among the tools commonly used for understanding the distribution of organic and inorganic nutrients within the complex soil matrix and their role in soil, as well as in total ecosystem stability and function, are a variety of conceptual and mathematical models.

Conceptual models of biogeochemical cycles are valuable for evaluating and testing our understanding of nutrient movement and distribution in an ecosystem. The degree of complexity of individual models necessarily varies with the objectives of the model developer. Rarely are all states (chemical as well as physical) of the nutrient pools represented in a specific model. Rather, those processes most important to understanding a particular ecosystem function are detailed. For example, a model of carbon cycling can be as simple as a depiction of the size of the carbon dioxide and organic carbon pools and assessment of the rates of movement of carbon between the two pools. Alternatively, movement of carbon between mineral (e.g. carbon dioxide and limestone) and soil organic matter, microbial biomass, and plant biomass pools may be evaluated in order to gain a better picture of the fate of soil organic components entering the soil system. Should specific organic compounds be of interest, the model may be concentrated on distributions and fate of specific chemical compounds (e.g. carbon dioxide, cellulose, lignin, and proteins).

The variability in the degree of complexity associated with development of biogeochemical cycles occurring in soil reflects the vast diversity of chemical and physical states of the elements (e.g. carbon, nitrogen, sulfur, phosphorus) that can be considered. Carbon is found as carbon dioxide in soil air, bicarbonate in soil water, and carbonate in soil minerals, not to mention the vast number of organic chemicals contained in living cells, soil humus, and any anthropogenic or xenobiotic compounds in the system. Added complexity is introduced when the elements of concerns exists in several oxidation states (e. g., nitrogen, sulfur as well as metal ions such as iron, and aluminum. Therefore, an evaluation of biogeochemical cycles must include consideration of the totality of biological, chemical, and physical interactions associated with the movement of nutrients between organic and inorganic pools and the variety of geochemical reservoirs existent within an ecosystem.

This definition of the complete realm of biogeochemical cycles emphasizes the point that study of these chemical transformations has real implications beyond the world of plant science and general biology. Study of these cycles of life‐sustaining nutrients is also essential for such broad‐based disciplines as general ecology or environmental science. For example, consideration of biogeochemical cycles in disturbed environments is essential for optimizing anthropogenic activities for ecosystem sustenance as well as in development of management plans for restoration or reclamation of damaged systems; that is, development of appropriate ecosystem stewardship procedures.

As suggested by the foregoing analysis, a study of biogeochemical cycles requires a means of assessing a wide array of complex interactions within a constantly varying environment. Some details may be studied in isolation, but it must be remembered that the primary objective of the exercise is to develop an understanding of the overall pattern of distribution and movement of nutrients in the ecosystem and their impact on ecosystem function and sustainability as well as evolution. A means of developing an overarching framework for study of biogeochemical cycles is to assemble the data into models. These models may be used to elucidate gaps in knowledge, provide hypotheses for future study, and develop experimental plans for quantifying nutrient movement within the cycles and assessment of the impact of environmental modification on these processes. Early modeling activity, or perhaps modeling activity at its simplest, involves conceptual models. The substances of interest are grouped in boxes (commonly referred to as pools) with arrows between the boxes to designate direction of nutrient flow. Some models include size and shape of the vectors of movement of the substances of interest between various chemical and physical pools existent in the site of interest. These vectors may involve processes of limited magnitude (e.g. decomposition of specific pesticides) as represented by major or minor pathways of nutrient flow (or in the example noted above, movement of the pesticide between various physical and chemical states of the chemical and the resultant changes in its impact of ecosystem function). These models provide a basis for design of experiments that yield the data necessary for development of more complex and predictive mathematical models.

Therefore, common objectives of study of nutrient flow in an ecosystem may include (i) the assessment of chemical forms of a nutrient element (e.g. carbon as carbon dioxide and organic matter; nitrogen as mineral and organic forms; phosphorus in soil minerals, water‐soluble, and organic forms) in an ecosystem of interest, (ii) the quantity of nutrients retained in these compounds (i.e. the size of the nutrient pool), and (iii) the rate of movement of the nutrients between these various nutrient pools as they affect the function and stability of various biological populations.

Once the participants of interest in biogeochemical cycles are determined (even in lieu of determination of all the pertinent chemical intermediates), conceptual models may be assembled, which are most easily presented as a series of boxes representative of the primary nutrient pools with arrows connecting the boxes to indicate the directional transformation (i.e. nutrient flux) of the cycle components (see Figure 9.1 for a depiction of the carbon cycle). To augment the information contained in a conceptual model, the type of arrow connecting the various reservoirs (e.g. dotted, size differential, or boldface) may be varied to suggest differences in the magnitude of the processes depicted. More commonly, a conceptual model is limited to indication of the nutrient reservoir and the direction of nutrient flow. Details such as distinction between relatively stable reservoirs, such as soil minerals or peat deposits, and the more ephemeral cycle components exemplified by simple organic compounds or mineral nitrogen forms may be minimal. That is, organic carbon may be considered as a single pool in the cycle or perhaps two pools, representing slowly and rapidly metabolized carbon compounds. A more quantitative representation of biogeochemical cycles or the individual reactions comprising them is entailed in extension of conceptual models into mathematical models as described in Section 9.1.2.

A generalized carbon cycle model, with four interconnecting boxes labeled “atmospheric CO2,” “humus microbial biomass,” “plant biomass,” “animal biomass,” and “reclaimed biomass.”

Figure 9.1 A generalized carbon cycle model with biological and atmospheric aspects emphasized.

A generalized model of the nitrogen cycle, with interconnecting boxes labeled “plant biomass,” “animal biomass,” “anthropogenically fixed nitrogen,” “atmospheric N2,” etc.

Figure 9.2 A generalized model of the nitrogen cycle.

Conceptual models of biogeochemical cycles emphasize the distribution of an element in a variety of chemical forms. The elements of interest are found to move between several chemical (organic vs inorganic) and/or oxidation states at predictable rates. Basic chemical components of plant and animal biomass are transformed from biomass unavailable forms to compounds that are readily incorporated into living cells. For example, the reasonably inert compound dinitrogen (N2) is reduced by soil microbes to ammonium, which, once it is released by the microbial cell, can be incorporated into plant tissue. With the death of the plant, the soil microbial community may return the nitrogen to the atmospheric pool of dinitrogen. This conceptualization of biogeochemical cycles is plant centered (or perhaps life centered, when microbial and animal biomass are also viewed as major nutrient recipients). Other models could be centered on movement of elements and compounds to and from atmospheric pools, soil minerals, humus, or any other primary ecosystem component, such as production and decomposition of greenhouse gases.

Conceptual models by their very design may provide a biased view of biogeochemical cycles in that most representations are limited to designation of the basic chemical and biological forms of the cycle components. To provide a more complete model for soil microbiological or environmental science considerations, designation of the path for movement of the elements of interest through various ecosystem components (i.e. soil minerals, atmospheric gases, and biotic entities) is useful. A sound ecological approach to study of biogeochemical cycles should include an understanding of:

· the interactions between the compounds transformed, the biological community catalyzing the processes, and the physical environment in which the reactions occur and which itself may be modified during the process

· the contribution of spontaneous chemical reactions to the overall nutrient generation rates.

In regard to specific transformations, the primary product of the physical, chemical, and biological processes comprising a biogeochemical cycle is a stable, sustainable ecosystem (that is, an ecosystem that can adapt to changes in the physical, chemical, and biological controllers of the soil biological community). Thus, although conceptual models of biogeochemical cycles are usually presented as “boxes connected by arrows,” it must be remembered that the boxes represent “real‐world” nutrient reservoirs, such as plant and microbial biomass. Furthermore, at least as important as the quantification of chemical form in a cycle is an understanding of the magnitude of the vectors representing the movement of the material of interest between the various nutrient pools. The importance of a large pool of nutrients (such as atmospheric N2), which in itself is unavailable to microbes, is determined not by its size but rather by the rate at which it is converted into a form useable by microbes (i.e. nitrogen fixation). The rate of the transformation is controlled not only by biological processes themselves but also by the properties of the environment in which the plant, microbe, or animal is located. Therefore, the occurrence of biogeochemical cycles is imperative for long‐term sustainability of the ecosystem.

Thus, collection of data for development of conceptual models should entail consideration of biotic as well as abiotic interactions. For example, mineral nitrogen (i.e. ammonium and nitrate) availability frequently controls the productivity and nature of the aboveground community. This nitrogen may be derived through nitrogen fixation or other sources external to the specific ecosystem under study or the nitrogen contained in biomass may be recycled into nascent biomass through mineralization of deceased biomass and soil organic matter pools. Thus, in this situation, the indigenous nitrogen mineralization rate is a primary controlling factor for biomass production. Note that in the latter situation, the total amount of nitrogen contained within the soil system may be reasonably constant. That is, the productivity of the aboveground community is limited at least in part by the capacity of the soil microbial community (nitrogen fixers or nitrogen mineralizers) to incorporate the soil nitrate and ammonium into plant biomass. The activity of this soil microbial community is in turn controlled by the quantities of energy and nutrients provided by the aboveground community in the form of photosynthate, as well as that contained in the soil physical and chemical environment. For example, in a high‐moisture, high‐clay soil, limitation of plant productivity by the physical properties of the system can also curtail soil microbial activity, which is dependent on decomposition of plant biomass for its energy source. Also, anoxic conditions may favor loss of fixed nitrogen from the system through biological denitrification reactions.

These life‐limiting environmental properties are continually being modified by the physical presence and biological products of the plant, animal, and microbial communities (see Tate 1987). For example, improvements in the soil physical environment through enhancement of soil aggregate structure by growth of the microbial community augments plant biomass production, which may further stimulate microbial growth through augmented inputs of fixed carbon.

Therefore, essential steps in the development and refinement of conceptual models involve elucidation of the forms of the nutrient element of interest in the system, perhaps tracing the movement of elements between these forms using such isotopes as 14C or 15N, and determination of the physical and chemical properties of the system controlling the relative magnitude of the various chemical pools and the rate of flux between these reservoirs (e.g. see Paterson et al. 2011; Tatzber et al. 2009) The process of model building is based primarily on characterization of the system and development of a reasonable understanding of the relationship between the various nutrient reservoirs occurring therein. These studies may entail determination of biological trophic interactions (e.g. Pennisi 2015; Santos et al. 1981, 1984; Santos and Whitford 1981), determination of the impact of management and climate change on ecosystem function (e.g. Allen 1990; Pastor and Post 1988; Tabatabai and Chae 1991; Walters et al. 1992), evaluation of the impact of properties of ecosystem inputs on their mineralization (e.g. Fresquez et al. 1990; Meentemeyer 1978), as well as examination of native ecosystem processes (e.g. Parker et al. 1983).

9.1.2 Mathematical Modeling of Biogeochemical Cycles

Common objectives of projects examining nutrient cycling in a soil ecosystem are twofold.

· Developing an understanding of the processes and their limitations in the particular site of interest. This type of experiment may involve achieving a clearer understanding of the nuances of nutrient availability in a native forest ecosystem, for example. Alternatively, these types of studies can provide the basis for nutrient management as exemplified in agricultural soils as well as for soil reclamation.

· Collection of a data pool that allows extension of the concepts learned at a few sites to a more extensive land area.

Conceptual and mathematical models aid in achievement of these types of objectives. A conceptual model provides a basis for the “design” of the mathematical representation of the biogeochemical cycles in an ecosystem. The components of the cycle and direction of flow of nutrients are described, but little is revealed about the size of the various pools or the magnitude of flow between the pools. Frequently, a conceptual model can represent a hypothesis of “how the system works,” thereby providing a basis for an experiment designed to test specific hypotheses. Initial steps for attaining an understanding of nutrient cycling processes include analysis of the quantities of various compounds present within the soil site of interest. Generally, it becomes clear that a detailed understanding of the intricacies of the reactions is obscured by site heterogeneity and the sheer number of potential soil properties that control the reaction rates and nutrient distribution. Thus, a mathematical approach to data summarization and interpretation is required.

First steps for developing more generally applicable mathematical relationships may stem from application of statistical analysis procedures. A simple model may be yielded through a linear or nonlinear regression analysis of the data. The resultant equations may be reasonably descriptive of the particular site under study, but applicability to other situations is limited. This conclusion stems from the fact that the equations derived are mathematical relationships fitted to aspecific data sets. That is, the relationship may be forced to fit a straight line, parabola, or hyperbola with the constants in the equations being representative of the conditions of the limited number of soil sites used in the experiment.

Equations more generally applied to biological, chemical, or physical reactions may be adapted to modeling environmental data. For example, an enzyme‐catalyzed reaction may be reasonably easily described by a first‐order equation or in some cases by a Michaelis–Menten relationship (see Chapter 6). As the complexity of the system increases (e.g. interest in the quantities of the reactant consumed within the profile of a conifer forest) and the variety of interactions with soil properties and climatic variations expands, inclusion of more general relationships, such as the Q10 equation or Ahrenius relationship for temperature impact on reaction rates, may be necessary to increase model utility. Furthermore, in the complex, heterogeneous environment of soil, effective chemical concentrations for a reaction may be defined by properties other than total substrate present. For example, reaction rates may be dependent upon soil parameters such as the rate of dissolution of the substrate (Stucki and Alexander 1987) or diffusion of the enzyme substrate into the microsite wherein the organisms reside (Scow and Alexander 1992; Scow and Hutson 1992). Also, in an ecosystem adapting to a perturbation, the rate of occurrence of a process may be proportional to the growth or death of the microbial populations producing the requisite enzymes (i.e. the reaction may be best described by a Monod equation – see Chapter 4).

Therefore, a variety of mathematical relationships descriptive of not only biological but also chemical and physical interactions are combined to provide an interactive series of equations descriptive of the extent of a natural process and predictive of rate changes with perturbation of the system. Because of the general applicability of the mathematical relationships (e.g. Michaelis–Menton equations provide a reasonable model of the kinetics of the vast majority of enzymatic processes), biogeochemical models based on them provide reasonable descriptors of process rates in a diversity of situations. It must be remembered that applicability of the mathematical relationship to alternative situations is still limited by the range of data used to validate the model. For example, a model descriptive of nitrogen transformation rates in well‐drained agricultural soils would have to be shown to be applicable to swamp or wetland soils were such soils of significant importances in the sites being studied. This situation results from the fact that the mathematical relationships used to describe soil processes are a mixture of variables and constants. Values for the variables are measured, estimated, or assumed from analysis of actual soil situations. The constants are based on or are commonly derived from these “real‐world” studies. Thus, equations and models are only applicable within the range of variation of the data used to develop (validate) them. Their applicability to other soils with properties outside that range must be proven. A simple mathematical model of soil nitrogen mineralization rates and organic nitrogen pool size, which is based on first‐order reaction kinetics, is described in Section 9.5.4.

Examples of the diversity of opportunities wherein studies of soil microbial interactions with the physical and chemical structure of their habit can benefit from development of simple to complex conceptual and mathematical models include the following.

· Assessment and elucidation of soil properties controlling specific portions of a biogeochemical cycle. Examples of such applications include study of the impact of distribution of aerobic and anaerobic microsites on denitrification rates (McConnaughey and Bouldin 1985a, 1985b; McConnaughey et al. 1985); sequestration of the chemical(s) of interest, and thereby reduction of availability of decomposable organic matter to soil microbes (see Dou et al. 2014; Wang and Sainju 2014), as well as evaluation of nitrogen mineralization (Deans et al. 1986), sulfur and nitrogen mineralization (Ellert and Bettany 1988) and carbon mineralization (Stroo et al. 1989) in a variety of ecosystems.

· Evaluation of ecosystem response to global changes such as determination of soil changes due to global warming (see Creamer et al. 2015; Jenkinson et al. 1991; Lal 2004; Luo et al. 2014).

· Quantification of specific fluxes between nutrient pools, for example, evaluation of the importance of nutrients contained in microbial biomass in nutrient cycling (Trumbore and Czimezik 2008; van Veen et al. 1984).

· Prediction of the impact of management changes on soil enzyme activity, microbial populations, and soil quality‐associated properties, such as nutrient levels and their distribution, variation in soil enzyme activities and microbial population distribution and function (Hunt 1977; Kennedy and Schillinger 2006; Parton et al. 1988; Paterson et al. 2011; Stott et al. 2013; Wallenius et al. 2011), or of erosion on plant phosphorus availability (Jones et al. 1984).

In each of these situations, varying combinations of mathematical relationships descriptive of biological transformations, pH variation, temperature effects on biological processes, adsorption processes, plant composition were combined to provide a mathematical representation of movement of nutrients between various biogeochemical nutrient pools.

9.2 Specific Models of Biogeochemical Cycles and Their Application

Basic conceptual models of the primary nutrient cycles (carbon, nitrogen, sulfur, and phosphorus) are depicted in Figures respectively. These models are designed to emphasize the essential components of the cycles. More obvious points to note are that the elements are transferred between organic and inorganic forms (all cycles) as well as through a variety of oxidation states (nitrogen and sulfur cycles). The oxidation state of nitrogen varies from −3 (ammonium ion) to +5 (nitrate ion) whereas in the sulfur cycle the oxidation state of the sulfur atom varies from −2 (sulfide) to +6 (sulfate). Similarly, a variety of soil and atmospheric inorganic compounds (e.g. nitrate, ammonium, dinitrogen, elemental sulfur, phosphate rock, sulfur‐bearing minerals) and plant, animal, and microbial biomass organic components constitute primary nutrient pools in each of the models. Note that geological materials are of major importance in both the sulfur and phosphorus cycles.

A model of the soil phosphorus cycle with soil mineral reservoirs as the central component, with interconnecting boxes labeled “soil organic – PO4,” “plant – PO4,” “animal – PO4,” “surface waters,” etc.

Figure 9.3 A model of the soil phosphorus cycle with soil mineral reservoirs as the central component.

A model of the sulfur cycle showing generalized interactions of biological resources with soil and atmospheric sources. Interconnecting boxes are labeled “plant biomass,” “animal biomass,” “atmospheric sulfur,” etc.

Figure 9.4 A model of the sulfur cycle showing generalized interactions of biological resources with soil and atmospheric sources.

Traditionally, biogeochemical cycles are studied in order to assess the relationship of nutrient state to higher plant productivity. This limited purview could lead to an underappreciation of the impact of nutrient cycling on total ecosystem functions and properties. A number of intermediates in the various cycles have considerable impact on the integrity of the terrestrial ecosystem. Positive as well as negative implications are discerned, especially from the viewpoint of societal impact on system function.

Perturbations of nutrient cycling could result in overproduction of substances that result in large‐scale declines in system quality, altered biomass productivity, or modified terrestrial ecosystem sustainability. A long‐appreciated example of this phenomenon is the changes in soil organic matter reserves due to soil cultivation (Tate 1987; Rosenzweiz and Hillel 1998) as well as total plant biomass productivity (e.g. Adams et al. 1990; Allen 1990; Dahlman et al. 1985; Reddy et al. 1989). Similarly, it has long been noted that soil organic nitrogen and carbon levels are generally enhanced when conservation tillage practices are implemented on soils previously managed with conventional tillage methods (e.g. Blevins et al. 1984; Cambardella and Elliott 1992; Dou et al. 2014; Doran 1980; Kennedy and Schillinger 2006). Loss of soil organic matter due to soil cultivation increases atmospheric carbon dioxide levels, thereby impacting global warming. In contrast, increased storage of carbon in soil organic matter pools moderates global warming due to greenhouse gas production. Furthermore, sulfate reduction to sulfide may be exploited to reduce the difficulties associated with acid mine drainage‐impacted soils, and denitrification may be stimulated to reduce surface and groundwater contamination with nitrate.

The ultimate impact of the soil microbial community on the total ecosystem is derived from its contribution to the orderly transition of the chemicals of life through their various oxidation states. Optimization of ecosystem conditions for nutrient cycling results in development of a sustainable system in which nutrients are cycled with minimal loss of essential components from the system or movement of potentially troublesome intermediates into portions of the ecosystem where their presence may be less appropriate (e.g. resulting in lake eutrophication or groundwater contamination). Microorganisms are instrumental in such essential ecosystem‐sustaining activities via the conversion of plant‐unavailable forms (e.g. N2) to chemical entities that may be incorporated into plant biomass (nitrate (NO3) or ammonium (NH4+)) as well as the return of organic components to inorganic forms, which may be recycled into plant biomass (carbon or nitrogen mineralization). Loss of plant biomass nutrients phosphate or nitrate negatively impacts the system by reducing its productivity and could be damaging to adjacent surface waters. Soil management to balance mineral nitrogen production and consumption minimizes surface‐ and groundwater quality decline.

9.2.1 The Environmental Connection

To truly understand the ramifications of any biogeochemical process, the implications of the synergism between the soil biota and their environment must be considered. Both the physical environment and the biological entities growing therein are modified by the products of biological growth. This synergism is absolutely essential for development of a sustainable ecosystem. In a managed site, such as a reclamation project, the sustainability of the developing ecosystem is reliant upon the ability of the microbial community to modify an initially hostile habitat to one more conducive for its own growth and sustainability as well as for the higher plant community. In highly damaged sites or in situations where rapid establishment of an esthetic aboveground community is desirable or required, anthropogenic intervention to manage microbially mediated processes may be necessary. In the absence of such site management, the longevity of the biological populations essential for establishment of a viable aboveground and belowground community could easily be truncated.

A typical scenario could be as follows. To provide a rapid start to the soil microbial community and to sustain aboveground populations until soil physical conditions are modified to those more conducive to plant development, the soils may be amended with an organic material (such as composts and sludges), nitrogenous fertilizers, and limestone to stimulate development of the indigenous microbial community. The organic amendments provide energy to the microbial communities as well as improving their growth conditions by complexing of heavy metals. The nitrogenous fertilizers are used to address imbalances in the carbon to nitrogen ratios of the organic components of the system. Liming of the soil may be necessary to adjust the soil pH to a range optimal for biological community development. Initially, this managed ecosystem development would be supported by the external nutrient inputs. A stable situation requires the slow evolution of soil populations and modification of soil physical properties to a point where nutrients required for continued biomass productions are produced internally. This transition from an artificially sustained to an internally sufficient system requires such physical‐biological synergistic interactions as enhancement of soil physical structure through aggregation (see Chapter 1), as well as microbial amelioration of soil acidity in sulfate‐impacted sites. This system development can be termed quasi‐synergistic in that a feedback interaction occurs between the developing microbial community and its physical environment. As the complexity of the microbial community increases, one product of its activity is an improvement in the physical and chemical conditions of its habitat. This alteration in turn supports further microbial growth. This cycle is repeated until ultimately a steady state develops with an optimized balance between the microbes and their physical and chemical environment.

These belowground processes not only optimize soil microbial community activity and development, but also improve conditions for plant population productivity and longevity. The development of soil aggregates through the action of soil microbes results in improved air diffusion and water infiltration rates. Therefore, not only is the soil microbial community stimulated by this habitat improvement but plant root growth and aboveground biomass synthesis are also increased. The response of the plant community to the microbially modified soil structure allows for further enhancement of soil microbial activity. The microbial community is limited, to a large degree, by the fixed carbon supply. This fixed carbon in nonmanaged ecosystems is the direct product of plant photosynthesis. Therefore, the improvement of soil structure noted above results in better conditions for plant growth, which further stimulates the microbial community through the enhanced input of photosynthate into the soil. Thus, the biological aspect of that portion of the carbon cycle represented by microbial mineralization has modified the physical environment so as to increase carbon fixation by the plants, which in turn allows for further enhancement of the soil microbial populations dependent on generation of this plant carbon. This portion of the ecosystem development could be considered to be a true synergistic interaction.

The interdependence of higher plants, microbes, and their physical environment (not simply from a fixed environment, but more from the view that these interactions are resulting in a continuously evolving ecosystem) is demonstrated by the simple example of site management noted above. Similar interactive effects can be exemplified by evaluating the biological and chemical interactions involved with acid mine drainage‐impacted soil. Many of the ores mined throughout the world are sulfide salts (e.g. iron sulfide and zinc sulfide). As long as the ecosystems containing the sulfide salts are not disturbed, the conversion of the sulfides to sulfates is limited by the anoxic or very low oxygen levels. Mining of the metal ores results in oxygenation of the metal sulfides to more water‐soluble metal sulfates. This process results in acidification of the soil sufficiently that the indigenous acid‐sensitive microbial population ceases to function. Reclamation of the soil requires raising the pH sufficiently that biological activity precluded by the acid may be resumed. Since biological sulfate reduction to sulfide requires anoxic conditions, site management must include flooding of the soil to encourage development of sulfate‐reducing bacteria. The ameliorative processes eventually result in reestablishment of biogeochemical cycling of nutrients. (See Mills (1985) for a further discussion of management of acid mine drainage.) The microbial role in this situation simply involves chemical reduction of the soluble sulfate salts to sulfide, i.e. the sulfate serves as a terminal electron acceptor for the microbial community, which results in precipitation of the metal sulfide under the anoxic conditions. Note that the anoxic conditions must be maintained to prevent the soil microbes from oxidizing the sulfide to sulfate as a metabolic energy source.

These examples demonstrate the optimization of microbial community development which results from modification of the physical and chemical environment due to the action of the biological components. Traditionally, this “optimization” is generally considered to entail improvements in community structure, which translates from an anthropogenic viewpoint into an esthetic aboveground community. However, ecosystem evolution may involve alterations in the chemical environment that cause declines in aboveground productivity. For example, drainage of acid peats (either artificially or anthropogenically induced) stimulates the microbially catalyzed oxidation of the metallic sulfides contained therein into sulfate. The resultant decline in soil pH results in death of acid‐sensitive plants in the aboveground community.

Biological cycling of nutrients in soils causes both permanent (long‐term) and transient (seasonal) alterations of the physical environment. Solubilization of minerals and their subsequent movement into plant tissue or even leaching from the soil site result in an irreversible change in soil mineralogy. In contrast, more ephemeral changes in soil components are associated with the interaction between the aboveground plant community and soil microbes. For example, total system fixed nitrogen loads may be increased through the activity of nitrogen‐fixing bacteria. This nutrient augmentation results in a general improvement in soil fertility. The effect may be limited if the newly incorporated fixed nitrogen is removed from the site (e.g. harvesting of crops, anthropogenically or by grazing animals), leached from the system in percolating waters, or denitrified.

The preceding analysis shows that both microbial and plant communities benefit from biogeochemical cycles. This gain by the biological community is enhanced by the associated modification of the physical environment by soil microbes. It must be stressed that no volition on the part of the communities or individual organisms is assumed in these adaptations. Beneficial alterations of the organisms’ surroundings are purely fortuitous. A situation is created wherein the microbial communities alter their habitat in a manner that results in selection of more effective microbial populations. This interaction produces a continuously evolving ecosystem.

9.2.2 Interconnectedness of Biogeochemical Cycle Processes

Commonly, the best way conceptualizing the biogeochemical cycles is to examine them as independent entities, as has been the historical educational practice. Yet, to some degree this presents an impediment to developing an absolute appreciation of the intricacies of nutrient cycling in the soil system. Central to all nutrient cycles are the various ecosystem components composed of organic matter – a variety of plant, animal, and microbial biomass pools plus intermediates in the mineralization of these components along with soil humic substances. This organic matter‐centered concept of biogeochemical cycling is depicted in Figure 9.5 (where the assembly of the components of biomass is shown) and Figure 9.6 representing the disassembly or mineralization of organic substances to the primary mineral pools.

Image described by caption and surrounding text.

Figure 9.5 Examples of the central role of plant biomass in sequestering various nutrient elements. (The subscript “o” refers to various organic pools of the individual element.)

The pivotal role of organic matter in biogeochemical cycles results primarily from the fact that it is the primary reservoir of carbon, oxygen, nitrogen, phosphorus, and sulfur in soil. Thus, in an unmanaged system, the movement of nutrients into this organic matter reservoir and their subsequent mineralization is the primary controller of plant nutrient availability. A conceptual model accentuating this fact is shown in Figure 9.7. The primary components of biomass can be envisioned as traveling along a highway that converges on a “square” consisting of living cells. Some organic substances are lost from the biomass and exit the square as biomass products (e.g. root exudates, extracellular enzymes), but most enter the “decomposition lane” upon demise of the cell. Some of the organic substances take a detour through the neighborhood of humic substances, but eventually all enter a square consisting of mixed composition organic compounds, e.g. amino acids, nucleotides, carbohydrates, phytin, and humic substances. Continued mineralization of these products yields the original mineral components.

Image described by caption and surrounding text.

Figure 9.6 Disassembly of plant organic matter into various primary inorganic nutrient pools. (The subscript “o” refers to various organic pools of the individual element.)

Image described by caption and surrounding text.

Figure 9.7 The road from an inorganic existence through biomass and return to inorganic forms. The subscript “i” refers to inorganic forms of the element.

A practical difficulty resulting from this interaction of the biogeochemical cycles is the designation of a particular mineralization process as being driven by the needs of an individual nutrient cycle. For example, amino acid mineralization in soil can be easily quantified. Major products of the process are cellular energy, carbon dioxide, ammonium, and sulfide. A question that could be considered is whether the ammonium is produced because of a requirement of the decomposer population for ammonium to synthesize new biomass or whether the mineral nitrogen was produced as a by‐product of the carbon and energy recovery processes. In fact, mineralization of carbon and nitrogen are sufficiently linked in soil that assessment of carbon dioxide evolution rates has been suggested as a means of estimating net nitrogen mineralization (Gilmour et al. 1985). The bottom line is that it is easy to produce models descriptive of various nutrient cycles, but difficult to designate particular processes as being primarily driven by needs of any specific nutrient cycle.

9.3 Biogeochemical Cycles as Sources of Plant Nutrients for Ecosystem Sustenance

A good way of gaining an appreciation of the interactive nature of the aboveground and belowground portions of the ecosystem is to examine the nutrient cycling interactions between the two regions. Nutrients enter soil through rainfall, dry deposition (e.g. sulfate, nitrate, and minerals contained in dusts), atmospheric gases (e.g. sulfur dioxide), anthropogenic sources (e.g. soil amendments), and fixation (dinitrogen, carbon dioxide). Primary sources of the basic building blocks of organic matter are divided between the atmosphere (carbon dioxide and dinitrogen) and soil (phosphate‐ or sulfate‐bearing soil minerals). Sulfur sources could be considered to be divided between soils (sulfur‐bearing minerals) and atmospheric sources (e.g. sulfur dioxide generated by power generation and, to some extent, automobile exhausts). With activities directed at alleviating air pollution, especially from automobiles and power plants, the latter source of sulfur compounds is being reduced, which has resulted in a need to consider soil mineral and organic matter derived sulfur inputs for plant biomass productivity.

As an ecosystem matures (approaches equilibrium), reliance on external nutrient inputs declines. The majority of the materials necessary for biomass productivity are produced by mineralization of resident organic matter pools. This situation is most clearly seen with nitrogen cycling in a mature forest. In a climax forest, the vast majority of the nitrogen is contained in soil organic matter and plant biomass nitrogen pools (e.g. Tate 1987). In these situations, nutrient availability to the plant community is reliant nearly totally on the rate at which the microbial community mineralizes plant, animal, and microbial biomass as well as soil humus.

Therefore, in ecosystems receiving minimal inputs of nutrients, a truncated nutrient cycle could be said to exist. The various nutrients are cycled between organic and mineral nutrient pools. Organic nutrient resources are composed of plant biomass and photosynthate entering the soil and native soil organic matter. These organic matter pools can be separated into (i) those substances delineated on the basis of solubility (humic acids, humin, and fulvic acids), and (ii) recognizable biochemical components of living cells; that is, polysaccharides, amino acids, nucleotides, etc. The latter substances are usually classed as readily available organic matter (see Chapter 10 for a discussion of the importance of this reservoir of organic material in ecosystem productivity). Readily available organic matter is that portion of the soil organic matter reservoir that is easily metabolized by the soil microbial community. It is distinguished by its short residence time (generally less than 1–2 years) and the fact that it provides essentially all the metabolic energy for the soil microbial community.

Ecosystems may also derive a major portion of their energy from externally supplied organic substances. These situations are usually the product of anthropogenic inputs. Specific examples are intensively cultivated agricultural soils, where crop debris and anthropogenically produced fertilizers provide the bulk of plant nutrients, plus soils impacted with varying inputs of sludges and composted materials. Soil amendment with these societal residues may be significant plant nutrient sources (e.g. see Boyle and Paul 1989; Geiger et al. 1992; Holland and Coleman 1987).

9.4 General Processes and Participants in Biogeochemical Cycles

The soil biological community mobilizes minerals, mineralizes organic substances, and immobilizes nutrients directly through a combination of oxidation and reduction reactions or indirectly through environmental modification. Direct reactions may involve alteration of the oxidation state of the particular element of interest as well as oxidation or reduction reactions. For example, energy recovery and mineralization of carbon‐based compounds involve oxidation (electron removal) of the carbon compounds to carbon dioxide. Similarly, ferrous iron may provide energy for microbial growth by removal of electrons to produce ferric iron. Alterations of the oxidation or reduction state are associated with oxidation of ammonium to nitrate and sulfide to sulfate. In each of these instances, the energy yielded by the oxidative electron transfer produces increases in microbial biomass.

Lest a false impression be created regarding the direction of the oxidation–reduction processes in soil, it should also be noted that a variety of reductive transformations are essential for integrity of the biogeochemical cycles. For example, whereas organic compounds are being oxidized by denitrifying bacteria, the electrons so generated are utilized to reduce nitrate to dinitrogen. Similarly, sulfate may also serve as an electron acceptor – producing hydrogen sulfide among the products – in accompaniment with oxidation of a variety of organic carbon compounds.

Minerals may also be solubilized within the ecosystem by less direct means than direct alteration of their redox state by actively growing microorganisms. For example, chelators solubilize metals and modification of the microsite pH may result in solubilization of a variety of cations. Acidification results in solubilization of metals such as aluminum as well as dissolution of carbonate rocks, thereby freeing associated cations.

The primary mediators of these processes are the soil bacterial, actinomycete, and fungal communities. Most research has historically been associated with these groups of organisms because they are responsible for catalyzing the bulk of the reactions involved in biogeochemical cycling. However, recall from Chapter 7 that the activity of these organisms in soil is controlled in part by their interactions with soil protozoa, nematodes, and microarthropods as well as a variety of soil animals (earthworms, ants, millipedes, and insects). Also, although the contributions of these latter biological groupings to nutrient cycling may be less than those of the primary decomposers, the higher members of the soil community also participate in mineralization of organic matter through the respiratory processes necessary to maintain their own biomass. For example, the interactions of protozoa and higher animals in soil are key to mineralization and cycling of nutrients through a variety of means.

· These organisms are instrumental in the mixing of organic matter within the soil profile, thereby augmenting the contact between the plant residues and the decomposer populations. Key participants in these processes include earthworms, ants, termites, and even burrowing animals to some extent.

· Soil animals are involved in inoculation of plant litter with propagules of decomposers. Movement of ants, termites, and other insects between soil nests and surface litter insures contact of active microbial cells with plant debris.

· The soil animals indirectly stimulate decomposition by modifying soil properties. For example, increased aeration results from formation of earthworm galleries, burrows and chambers, and anthills.

· The soil animals may increase the surface area of the organic debris indirectly through their movement through dry deposits of plant biomass as well as directly via chewing and tearing of plant parts. Increased surface area augments potential contact with decomposer populations.

· As indicated above, the organisms may mineralize the plant biomass directly. For example, the woodlouse (Porcello scaber) feeds on decaying pine needles (Soma and Saito 1983).

· Trophic interactions as described in Chapter 7 also stimulate decomposition. Interactions between the primary decomposers (bacteria and fungi) with population‐controlling nematodes, protozoa, and mites are key in maintaining active primary decomposer populations.

In summary, all soil decomposer populations, primary plus secondary decomposers, derive energy and reducing power (i.e. everything necessary for formation of a stable community) from the transformation of organic and mineral intermediates of the various nutrient cycles. Indirectly, the products of the decomposer populations benefit aboveground populations – plants directly and animals indirectly.

9.5 Measurement of Biogeochemical Processes: What Data Are Useful?

Design of field and laboratory experiments for the assessment of biogeochemical processes is complicated by the generally prevalent desire to extrapolate the data from the level of the soil sample contained within a test tube to an ecosystem‐wide or even whole‐Earth view. Questions impinging on validity of data extrapolations include considerations of the inclusiveness (or appropriateness) of the data collected and the selection of representative soil samples. Furthermore, manipulation of the soil sample (i.e. storage conditions, mixing, air drying) subsequent to collection may preserve or alter the relationship of process rates occurring in the sample from those occurring in the field.

9.5.1 Assessment of Biological Activities Associated with Biogeochemical Cycling

As was introduced in Chapter 2, a variety of methods are available for assessing biological activity in soil samples. Furthermore, a multitude of assays of enzymatic activities, which are useful for the estimation of transformation rates associated with various biogeochemical processes (e.g. dehydrogenase, phosphatase, as well as assessment of the oxidation rates of a variety of carbonaceous substrates, such as glucose and aromatic compounds), have been adapted for use with soil samples. Applicability of the laboratory assessments of soil systems to the field situation is controlled by inherent limitations of the procedures themselves as well as by difficulties associated with provision of a meaningful extrapolation of the results beyond the individual soil sample studied.

Problems with quantifying biological activities in soil for determination of nutrient cycling potential and rates can be exemplified by evaluating applicability of various measures of soil microbial biomass. Several methods have been described for estimating microbial biomass levels and activity (e.g. respiratory activity [oxygen uptake or carbon dioxide evolution] as well as dehydrogenase activity, chloroform fumigation, ATP measurement, PCR analysis, and direct examination). See Chapter 2 for a discussion of each of the methods. Considerations associated with utilization of each procedure involve the intrinsic limitations of the assay (i.e. what is and is not measured) plus variability of microbial biomass densities between soil samples and within a soil site.

For any scientific procedure, the researcher must specify what the assay is truly measuring. For example, it is easy to postulate that carbon dioxide evolution rates from soil are proportional to the activity of the decomposer population in that soil sample, without realizing that carbon dioxide is also produced by those organisms feeding on the primary decomposers plus any plant root tissue in the sample. Thus, such data likely yield an overestimation of carbon mineralization in an ecosystem. Similarly, it is reasonable to consider that the total active soil microbial biomass would be related to carbon mineralization potential in a soil sample, but microbial biomass assessment techniques – chloroform fumigation, ATP measurement and direct counts – all result in the quantification of both active and inactive microbial propagules.

9.5.2 Soil Sampling Aspects of Assessment of Biogeochemical Cycling Rates

Data interpretation is also complicated by the fact that the assays are generally conducted in the laboratory. Assuming that a representative field soil sample may be collected, data from analysis of the sample in the laboratory may not reflect actual population densities or activities in the field. A variety of changes in the soil properties generated by removal of the sample from its native site, its storage conditions, and preparation for study may result in alteration of the biological activity occurring therein. For example, collection of the soil sample as well as any sieving or mixing of the soil that occurs in the laboratory releases native soil organic matter trapped within soil aggregates. At least a portion of this organic matter can be mineralized by the microbial community. Thus, the biomass and the enzymes required for catabolism of the organic matter may be augmented above levels previously existing in the field. This complication is increased if storage of the sample is necessary. Even at refrigerator temperatures, significant growth of microbial populations can occur. Some enzymatic activities are preserved at reduced temperatures, whereas other activities are not (see Tate 1987 for a more complete discussion of this topic). Therefore, for any measure of biological activity in a field sample, control experiments must be conducted to determine stability of the variable of interest during storage.

Extrapolation of laboratory‐generated data to the field from which the sample was collected is reliant upon the sample being representative of the field situation. This prerequisite for data analysis is nowhere more essential than in studies of biogeochemical cycles. With few exceptions, laboratory assays are conducted with soil sample sizes ranging from less than 1 g to perhaps 100 g at most. As indicated in Chapter 1, horizontal and vertical variation of microbial populations in soils as well as the containment of microbial populations within discrete microsites within the soil matrix makes collection of manageable soil samples that are representative of an entire field or ecosystem site difficult. To produce data in the laboratory that are at least partially representative of the field situation, care must be taken to collect a series of soil samples that represent pertinent portions of the ecosystem of interest. This representative soil sample must then be preserved in a manner to minimize or preclude alteration of biological activities prior to their quantification.

9.5.3 Environmental Impact of Nutrient Cycles

As indicated above, biogeochemical cycles are composed of a variety of chemical compounds, linked in a cyclic manner, whose interconversion is accomplished by a combination of biological and chemical processes. Although a mass of data has been published documenting the size of various nutrient pools within a number of ecosystems, considerable interest must be directed at the rate of movement between these pools and the environmental impact of these transformations. Alteration of the distribution within the various pools or their rate of change will necessarily alter the ecosystem and adjacent system function. Frequently, complications arise from the failure to fully consider inputs or losses from the soil system. For example, nutrients may leave a system through leaching, denitrification, volatilization (including ammonium volatilization) as well as being lost in harvested crop. The latter category is not limited to that biomass removed through agricultural processes. Wild and domestic animals may be prime movers of nutrients from an ecosystem through their grazing activities. Positive as well as negative system impacts are easily encountered.

This movement of nutrients between various reservoirs has been exploited for agricultural crop production. “Mining” of fixed nitrogen contained in soil humus has long been the basis for successful subsistence agriculture, but this enhancement of nutrient transfer from organic to mineral reservoirs is not limited to low‐input, limited harvest cropping systems. Histosols, organic soils, are also so exploited. Peats and swamps represent sidetracks in nutrient cycles. Carbon, nitrogen, phosphorus, and sulfur compounds are incorporated into plant biomass which due to waterlogged conditions may be preserved for thousands of years. Removal of impediments to microbial mineralization of the organic matter results in rapid mineralization of the biomass. This produces sufficient nitrogen to support an intensive agricultural cropping system, as exemplified by the Everglades Agricultural Area (South Florida). Decomposition of the organic matter comprising the muck soils of that region produces as much as 3.3 × 104 kg C ha−1 year−1. This is accompanied by production of approximately 1400 kg N ha−1 year−1 (Tate 1980). For comparison, soils of a tropical rainforest are estimated to produce 405–2117 g C m2 year−1 (Schlesinger 1977). The above carbon evolution rate for the histosols in southern Florida is equivalent to 3300 g C m2 year−1.

In undisturbed ecosystems, such as mature forests, mineralization and biomass synthesis rates are nearly balanced. That is, recovery of the mineral nitrogen produced by the soil microbial community through plant roots is sufficiently efficient that little nitrogen leaves the site through such processes as leaching to groundwater. These slight losses due to percolation of mineral nutrients to ground or through surface water flow are balanced by nitrogen inputs due to nitrogen fixation. Concerns arise from those systems, especially disturbed sites, where nitrogen leaching may reach ground or surface waters (such as the movement of nitrate and phosphates into Lake Okeechobee due to oxidation of the histosols of the Everglades Agricultural Area). Economic considerations arise when externally supplied nutrients (fertilizers) added to agricultural soils leach from the system rather being retained in the crop.

9.5.4 Example of Complications in Assessing Soil Nutrient Cycling: Nitrogen Mineralization

Nitrogen mineralization rate is a key indicator of the potential biomass production that can be supported by indigenous mineral nitrogen production. As indicated above, the vast majority of the nitrogen contained in soils exists as organic nitrogen. Thus, in the absence of exogenous inputs, new biomass cannot be synthesized without mineralization of the resident organic nitrogen pools. The size of these nitrogen pools (organic and mineral) and flux rate between them are therefore meaningful values in characterization of an ecosystem.

Two procedures most commonly used to measure soil nitrogen mineralization potential are (i) the Stanford and Smith (1972) procedure, which was originally proposed for the study of soils contained in a column but has also been used for soils incubated as a batch sample in a beaker, plus (ii) the buried bag method (Eno 1960; Pastor et al. 1984). With the Stanford and Smith method, soil samples are collected and transported to the laboratory where they are sieved to increase homogeneity and to remove plant debris. The prepared soil samples are then dispensed into soil columns (leached method) or distributed into beakers (batch method). After an appropriate incubation period, mineral nitrogen is extracted from these samples with a dilute CaCl2 solution. Nitrate plus nitrite and ammonium production are assayed over a time frame usually varying from two to eight weeks. With the buried bag procedure, soils are collected in the field and placed either with minimal manipulation or following sieving into a polyethylene bag. The bag is returned to the soil horizon in a position as closely replicating the situation of the collected soil as possible. In theory, the conditions within the bag represent the field situation except that free exchange of soil water is prevented. That is, it is usually assumed that gaseous conditions within and without the bag are equivalent. Data from Bremner's laboratory (Bremner and Douglas 1971) indicate that free oxygen concentrations may decline inside the bag to levels that would reduce the activity of aerobic microbial populations. For an example of the use of these two procedures with forest soils, see Poovarodom and Tate (1988) and Poovarodom et al. (1988). With each procedure, mineral nitrogen production rates are determined by quantifying ammonium and nitrate plus nitrite concentrations in the soil samples before and following incubation.

The primary objective of conducting these assays is to measure the quantities of organic nitrogen that can be mineralized (size of the nutrient pool) and how fast it is being mineralized (flux rate). Note that these two values are not directly measured by the above procedures. Although direct assessment of the size and mineralization rate of labile soil organic nitrogen pool would be ideal, due to the complexity of the soil organic nitrogen pool, this is impossible. Soil organic nitrogen compounds are a highly diverse mixture of compounds, only a portion of which is easily metabolized by the soil microbial community. Furthermore, no soil organic nitrogen component has been found to be directly proportional to the mineralization rates. Therefore, these values must be derived indirectly. Both the quantities of mineralizable nitrogen contained within a soil sample and its mineralization rate are calculated from nitrogen mineralization data. Such data are predicated upon the assumptions that (i) soil organic nitrogen mineralization occurs in the laboratory‐incubated or buried bag samples at rates comparable to the field situation, (ii) preparation of the soil samples for analysis does not alter the availability of the organic nitrogen to the microbial community, and (iii) mineralization is a first‐order process occurring at the same rate throughout the incubation period.

Determination of the nitrogen mineralization rate (Stanford and Smith procedure as well as buried bag technique) is usually predicated upon the observation that nitrogen mineralization follows first‐order kinetics:

(9.1)equation

where N is the concentration of mineralizable organic nitrogen, k is a rate constant and t is time. Integration of this equation produces the following:

(9.2)equation

where Nt is organic nitrogen at time t and NO is organic nitrogen at time zero. Since the nitrogen at time t is equivalent to the initial nitrogen minus the nitrogen mineralized (Nm), the equation can be reduced to one containing a single unknown:

(9.3)equation

(9.4)equation

(9.5)equation

These calculations provide an estimate of the original organic nitrogen and its mineralization rates under the conditions where the data were collected (i.e. the laboratory or within the buried bag). Thus, the derived parameters are not necessarily applicable to true field situations. For example, NO represents the initial nitrogen concentration in the disturbed soil sample; that is, the nitrogen available to the microbial population after the aggregates were disrupted by sampling and sieving of the soil sample. A significant proportion of that nitrogen could have been trapped in soil aggregates prior to sampling. Thus, the mineralization rate derived from the laboratory‐incubated and buried bag retained soils may be an overestimation of that occurring in the field.

Further inaccuracies in the data derived by this method are added by the potential variation in nitrogen production due to diminished oxygen availability in the buried bag procedure (and the potential for this to occur in batch‐incubated samples), which would result in an underestimation of Nm. Complications associated with the Stanford and Smith leaching procedure involve the removal of mineralizable organic matter during leaching of the columns (Smith et al. 1980). With the latter nitrogen mineralization assay method, the soil columns are leached with a salt solution, prior to and at intervals during the incubation time. The resultant leachate contains the soil mineral nitrogen, which is the objective of the leaching procedure. It also includes a portion of the water‐soluble organic nitrogen. Removal of this latter soil nitrogen fraction from the pool of organic nitrogen may reduce the actual mineralization observed during the incubation procedure. Each of these difficulties in incubation of the soils – modification of soil atmosphere and removal of mineralizable organic nitrogen – may alter the rate of ammonium, nitrate, and nitrite production by the soil microbial community. Thus with all of these procedures, at best, an estimate of the field‐available organic nitrogen pool and its mineralization rate is provided.

The model described above is based on observations that nitrogen mineralization in soil generally follows first‐order kinetics. Other mathematical models for deriving NO and k have been proposed that are based on alternative nitrogen mineralization kinetics (e.g. Bonde and Lindberg 1988; Broadbent 1986; Cabrera and Kissel 1988; Deans et al. 1986; Juma et al. 1984), but the first‐order model described above appears to be the most commonly used for derivation of these nitrogen mineralization constants.

These observations demonstrate the degree of difficulties that are encountered in producing data representative of field nutrient pool size and the flux between these various pools. Such considerations of cycles where large mineral pools exist are complicated by difficulties in assessing small changes in mineral nutrient against a large background. This problem is commonly encountered with measurement of fluxes of organic phosphate and sulfate against backgrounds of soil‐soluble and rock phosphate. These observations clearly reveal the basis for problems with ecosystem‐wide and global extrapolation of such data.

9.6 Conclusions

An appreciation of the properties of the reactions occurring in biogeochemical cycles and the impact of variation in environmental factors on their reaction rates is essential for a complete understanding of a native ecosystem and in development of soil management or reclamation plans. These cycles are commonly appreciated solely for their contributions to aboveground plant productivity and their impact on ecosystem development, but it must be understood that biogeochemical processes are much more complex than the simple biological or chemical catalyzed transformation of an essential plant macro‐ or micronutrient. A complete depiction of biogeochemical cycles must entail an understanding of not only the biological entities involved but also their synergistic interactions with the soil physical and chemical properties as they contribute to development of the native ecosystem. For reclaimed soils, this synergism impacts development of a self‐sustainable ecosystem.

References

1. Adams, R.M., Rosenzweig, C., Pert, R.M. et al. (1990). Global climate change and US agriculture. Nature 345: 219–224.

2. Allen, L.H. Jr. (1990). Plant responses to rising carbon dioxide and potential interactions with air pollutants. J. Environ. Qual. 19: 15–34.

3. Blevins, R.L., Smith, M.S., and Thomas, G.W. (1984). Changes in soil properties under no‐tillage. In: No‐Tillage Agriculture (eds. R.E. Phillips and S.H. Phillips), 190–230. New York: Van Nostrand Reinhold.

4. Bonde, T.A. and Lindberg, T. (1988). Nitrogen mineralization kinetics in soil during long‐term aerobic laboratory incubations: a case study. J. Environ. Qual. 17: 414–417.

5. Boyle, M. and Paul, E.A. (1989). Nitrogen transformations in soils previously amended with sewage sludge. Soil Sci. Soc. Am. J. 53: 740–744.

6. Bremner, J.M. and Douglas, L.A. (1971). Use of plastic films for aeration in soil incubation experiments. Soil Biol. Biochem. 3: 289–296.

7. Broadbent, F.E. (1986). Empirical modeling of soil nitrogen mineralization. Soil Sci. 141: 208–213.

8. Cabrera, M.L. and Kissel, D.E. (1988). Evaluation of a method to predict nitrogen mineralization from soil organic matter under field conditions. Soil Sci. Soc. Am. J. 52: 1027–1031.

9. Cambardella, C.A. and Elliott, E.T. (1992). Particulate soil organic‐matter changes across a grassland cultivation sequence. Soil Sci. Soc. Am. J. 56: 777–783.

10. Creamer, C.A., de Menezes, A.B., Krull, E.S. et al. (2015). Microbial community structure mediates response of soil C decomposition to litter addition and warming. Soil Biol. Biochem. 80: 175–188.

11. Dahlman, R.C., Strain, B.R., and Rogers, H.H. (1985). Research on the response of vegetation to elevated atmospheric carbon dioxide. J. Environ. Qual. 14: 1–8.

12. Deans, J.R., Molina, J.A.E., and Clapp, C.E. (1986). Models for predicting potentially mineralizable N and decomposition rate constants. Soil Sci. Soc. Am. J. 50: 323–326.

13. Doran, J.W. (1980). Soil microbial and biochemical changes associated with reduced tillage. Soil Sci. Soc. Am. J. 44: 765–771.

14. Dou, F., Hons, F.M., Wright, A.L. et al. (2014). Soil carbon sequestration in sorghum cropping systems: evidence from stable isotopes and aggregate‐size fractionation. Soil Sci. 179: 68–74.

15. Ellert, B.H. and Bettany, J.R. (1988). Comparison of kinetic models for describing net sulfur and nitrogen mineralization. Soil Sci. Soc. Am. J. 52: 1692–1702.

16. Eno, C.F. (1960). Nitrate production in the field by incubation of soil in polyethylene bags. Soil Sci. Soc. Am. J. 24: 277–279.

17. Fresquez, P.R., Francis, R.E., and Dennis, G.L. (1990). Sewage sludge effects on soil and plant quality in a degraded, semiarid grassland. J. Environ. Qual. 19: 324–329.

18. Geiger, S.C., Manu, A., and Bationo, A. (1992). Changes in a sandy Sahelian soil following crop residue and fertilizer additions. Soil Sci. Soc. Am. J. 56: 172–177.

19. Gilmour, J.T., Clark, M.D., and Sigua, G.C. (1985). Estimating net nitrogen mineralization from carbon dioxide evolution. Soil Sci. Soc. Am. J. 49: 1398–1402.

20. Holland, E.A. and Coleman, D.C. (1987). Litter placement effects on microbial and organic matter dynamics in an agroecosystem. Ecology 68: 425–433.

21. Hunt, H.W. (1977). A simulation model for decomposition in grasslands. Ecology 58: 469–484.

22. Jenkinson, D.S., Adams, D.E., and Wild, A. (1991). Model estimates of CO2 emissions from soil in response to global warming. Nature 351: 304–306.

23. Jones, C.A., Cole, C.V., Sharpley, A.N., and Williams, J.R. (1984). A simplified soil and plant phosphorus model: I. documentation. Soil Sci. Soc. Am. J. 48: 800–805.

24. Juma, N.G., Paul, E.A., and Mary, B. (1984). Kinetic analysis of net nitrogen mineralization in soil. Soil Sci. Soc. Am. J. 48: 753–757.

25. Kennedy, A.C. and Schillinger, W.F. (2006). Soil quality and water intake in traditional‐till vs. no‐till paired farms in Washington's Palouse region. Soil Sci. Soc. Am. J. 70: 940–949.

26. Lal, R. (2004). Soil carbon sequestration impacts on global climate change and food security. Science 304: 1623–1626.

27. Luo, C., Rodriguez‐R, L.M., Johnston, E.R. et al. (2014). Soil microbial community responses to a decade of warming as revealed by comparative metagenomics. Appl. Environ. Microbiol. 80: 1777–1786.

28. McConnaughey, P.K. and Bouldin, D.R. (1985a). Transient microsite models of denitrification: I. Model development. Soil Sci. Soc. Am. J. 49: 886–891.

29. McConnaughey, P.K. and Bouldin, D.R. (1985b). Transient microsite models of denitrification. II. Model results. Soil Sci. Soc. Am. J. 49: 891–895.

30. McConnaughey, P.K., Bouldin, D.R., and Duxbury, J.M. (1985). Transient microsite models of denitrification: III. Comparison of experimental and model results. Soil Sci. Soc. Am. J. 49: 896–901.

31. Meentemeyer, V. (1978). Macroclimate and lignin control of litter decomposition rates. Ecology 59: 465–472.

32. Mills, A. (1985). Acid mine waste drainage: microbial impact on the recovery of soil and water ecosystems. In: Soil Reclamation Processes. Microbiological Analyses and Applications (eds. R.L. Tate III and D.A. Klein), 35–81. New York: Marcel Dekker.

33. Parker, L.W., Miller, J., Steinberger, Y., and Whitford, W.G. (1983). Soil respiration in a Chihuahuan desert rangeland. Soil Biol. Biochem. 15: 303–309.

34. Parton, W.J., Stewart, J.W.B., and Cole, C.V. (1988). Dynamics of C, N, P and S in grassland soils: a model. Biogeochemistry 5: 109–131.

35. Pastor, J. and Post, W.M. (1988). Response of northern forests to CO2 induced climate change. Nature 334: 55–58.

36. Pastor, J., Aber, J.D., and McClaugherty, C.A. (1984). Above‐ground production and N and P cycling along a nitrogen mineralization gradient in Black Hawk Island, Wisconsin. Ecology 65: 256–268.

37. Paterson, E., Neilson, R., Midwood, A.J. et al. (2011). Altered food web structure and C‐flux pathways associated with mineralisation of organic amendments of agricultural soil. Appl. Soil Ecol. 48: 107–116.

38. Pennisi, E. (2015). Africa's soil engineers: termites. Science 347: 596–597.

39. Poovarodom, S. and Tate, R.L. III (1988). Nitrogen mineralization rates of the acidic, xeric soils of the New Jersey Pinelands: laboratory studies. Soil Sci. 145: 337–344.

40. Poovarodom, S., Tate, R.L. III, and Bloom, R.A. (1988). Nitrogen mineralization rates of the acidic, xeric soils of the New Jersey Pinelands: field rates. Soil Sci. 145: 257–263.

41. Reddy, V.R., Baker, D.N., and McKinion, J.M. (1989). Analysis of effects of atmospheric carbon dioxide and ozone on cotton yield trends. J. Environ. Qual. 18: 427–432.

42. Rosenzweiz, C. and Hillel, D. (1998). Climate Change and the Global Harvest. New York: Oxford University Press.

43. Santos, P.F. and Whitford, W.G. (1981). The effects of microarthropods on litter decomposition in a Chihuahuan desert ecosystem. Ecology 62: 654–663.

44. Santos, P.F., Phillips, J., and Whitford, W.G. (1981). The role of mites and nematodes in early stages of buried litter decomposition in a desert. Ecology 62: 664–669.

45. Santos, P.F., Elkins, N.Z., Steinberger, Y., and Whitford, W.G. (1984). A comparison of surface and buried Larrea tridentata leaf litter decomposition in North American hot deserts. Ecology 65: 278–284.

46. Schlesinger, W. (1977). Carbon balance in terrestrial detritus. Annu. Rev. Ecol. Syst. 8: 51–81.

47. Scow, K.M. and Alexander, M. (1992). Effect of diffusion on kinetics of biodegradation: experimental results with synthetic aggregates. Soil Sci. Soc. Am. J. 56: 128–134.

48. Scow, K.M. and Hutson, J. (1992). Effect of diffusion and sorption on kinetics of biodegradation: theoretical considerations. Soil Sci. Soc. Am. J. 56: 119–127.

49. Smith, J.L., Schnabel, R.R., McNeal, B.L., and Campbell, G.S. (1980). Potential errors in the first‐order model for estimating soil nitrogen mineralization potentials. Soil Sci. Soc. Am. J. 44: 996–1000.

50. Soma, K. and Saito, T. (1983). Ecological studies of soil organisms with reference to the decomposition of pine needles. II. Litter feeding and breakdown by the woodlouse, Porcellio scaber. Plant Soil 75: 139–151.

51. Stanford, G. and Smith, J.J. (1972). Nitrogen mineralization potentials in soils. Soil Sci. Soc. Am. J. 36: 465–472.

52. Stott, D.E., Karlen, D.L., Cambardella, C.A., and Harmel, R.D. (2013). A soil quality and metabolic activity assessment after fifty‐seven years of agricultural management. Soil Sci. Soc. Am. J. 77: 903–913.

53. Stroo, H.F., Bristow, K.L., Elliott, L.F. et al. (1989). Predicting rates of wheat residue decomposition. Soil Sci. Soc. Am. J. 53: 91–99.

54. Stucki, G. and Alexander, M. (1987). Role of dissolution rate and solubility in Biodegradation of aromatic compounds. Appl. Environ. Microbiol. 53: 292–297.

55. Tabatabai, M.A. and Chae, Y.M. (1991). Mineralization of sulfur in soils amended with organic wastes. J. Environ. Qual. 20: 684–690.

56. Tate, R.L. III (1980). Microbial oxidation of organic matter of histosols. In: Advances in Microbial Ecology, vol. 4 (ed. M. Alexander), 169–201. New York: Plenum.

57. Tate, R.L. III (1987). Soil Organic Matter. Biological and Ecological Effects. New York: Wiley.

58. Tatzber, M., Stemmer, M., Spiegel, H. et al. (2009). Decomposition of carbon‐14‐labeled organic amendments and humic acids in a long‐term field experiment. Soil Sci. Soc. Am. J. 73: 744–750.

59. Trumbore, S.E. and Czimezik, C.I. (2008). An uncertain future for soil carbon. Science 321: 1455–1456.

60. van Veen, J.A., Ladd, J.N., and Frissel, M.J. (1984). Modelling C and N turnover through microbial biomass in soil. Plant Soil 76: 257–274.

61. Wallenius, K., Rita, H., Mikkonen, A. et al. (2011). Effects of land use on the level, variation and spatial structure of soil enzyme activities and bacterial communities. Soil Biol. Biochem. 43: 1464–1473.

62. Walters, D.T., Aulakh, M.S., and Doran, J.W. (1992). Effects of soil aeration, legume residue, and soil texture on transformations of macro‐ and micronutrients in soils. Soil Sci. 153: 100–107.

63. Wang, J.J. and Sainju, U.M. (2014). Aggregate‐associated carbon and nitrogen affected by residue placement, crop species, and nitrogen fertilization. Soil Sci. 179: 153–165.

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