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2 Conceptual Models: How Factors, Processes, and Levels Regulate N Gas Emission from Soil Hole-In-Pipe Model

2 Conceptual Models: How Factors, Processes, and Levels Regulate N Gas Emission from Soil Hole-In-Pipe Model

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48



Ed Gregorich et al.



Figure 5 Diagram of hole-in-the-pipe conceptual model. Adapted from Davidson

(2000).



3.2.1 N Gas Emission as Function of Soil Pore Space Properties

The production and emission of N gases from soil are primarily the result of

biological activity, but that N-gas generating activity can be regulated to a

large extent by soil structural properties (Ball, 2013). The production and

emission of N2O from soil are regulated by three main factors: substrate

availability, aeration status, and temperature (Smith, 1980). These three

factors interact and are affected by the timing of weather events and soil

physical conditions.

Linn and Doran (1984) proposed a simple model, which illustrates the

extent to which water content, expressed as water-filled pore space, controls

the production of N2O and CO2. In their model, the critical point at which

dominant aerobic/anaerobic processes begin to exert control on the production of these gases is at 60% water-filled pore space, which approximates

field capacity for most soils with a loam texture. Field capacity is defined

as the soil water content after excess water has been drained from the soil;

at this water content the soil macropores have drained and are air-filled

but the micropores are still water-filled. This breakpoint of 60% water-filled

pore space is assumed to be the point of transition at which oxidative and

reductive processes are active.

This model is useful for a general understanding of the effects of soil

structure, as characterized by water-filled pore space, on N gas emissions



Nitrogenous Gas Emissions from Soils and Greenhouse Gas Effects



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from soil. A more detailed understanding can be obtained by evaluating the

size and distribution of soil pores. When a soil is compacted, for example, the

volume of pores is reduced. Not all pores are reduced similarly however,

because the largest pores are usually lost or compressed first (Richard

et al., 2001). This preferential loss of larger pores changes the pore size distribution, thereby affecting relative diffusivity, air permeability, and oxygen

content (Ball, 2013), which in turn control the production and emission of

N2O. Thus, the increased denitrification commonly observed in compacted

soils may not be attributed only to a reduction in overall soil pore space but

more immediately by preferential loss of large pore spaces tied to diffusive

transport of soil gas, resulting in suboxic or anoxic conditions in soil microsites (Smith et al., 2003; Gregorich et al., 2014).

3.2.2 N Gas Emission as Function of Temperature

Like soil water content, temperature affects N2O production and consumption both directly and indirectly. The direct effect of temperature is on the

kinetics of enzyme activity (Billings and Tiemann, 2014), but temperature

also indirectly affects N2O production by influencing the number, type,

and activity of nitrifying and denitrifying N2O-producing microorganisms

as well as the physiological processes performed by these organisms. Furthermore, enhanced microbial activity occurring as a result of higher temperatures leads to increased O2 consumption during respiration creating

microaerophilic or even anaerobic conditions that favor denitrification

and increase N2O production.



4. ROLE OF AGRICULTURAL MANAGEMENT

PRACTICES

Every facet of agricultural management has an impact on soil physical,

chemical, and biological properties and processes and thus can influence

nitrogenous gas emissions from soils (Figure 6) and the resultant greenhouse

gas effects. Addition of fertilizers and manure to supply N and organic matter

for crop growth is one of the most important management practices that

affect N emission because it supplies mineral N, the basic substrate for nitrification and denitrification processes. The type of inorganic or organic fertilizer, the rate at which it is applied, the time of application and where it is

placed in or on the soil all affect the kinetics of production and emission of N

gases. Tillage affects the soil structure in multiple ways (e.g., aggregation

processes and pore space properties) and these structural changes in turn



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Ed Gregorich et al.



Figure 6 Factors influencing direct and indirect N2O emissions from cropping systems.

Adapted from Venterea et al. (2012).



affect the amount of oxygen and N gases and their diffusive transport

through the profile to the soil surface. The type of crop grown, its sequence

in the crop rotation, and the amount of crop residues returned to the soil are

important because they affect the release and availability of available C and

N (Aulakh et al., 2001; Baggs et al., 2003).



4.1 Synthetic Fertilizers

4.1.1 Synthetic FertilizersdRate and Type or Formulation

The rate and type of synthetic fertilizers both have an effect on N gas emissions from soil. Nitrous oxide emission is often related exponentially to N

fertilizer application rate (Chantigny et al., 1998; Ma et al., 2010), but methodologies proposed by the IPCC (Smith et al., 2007) assume a linear

response, typically amounting to 1e1.2% of the N applied (Bouwman,

1996; Helgason et al., 2005; Gregorich et al., 2005; Mosier et al., 2006).

This loss rate is highly variable and can vary by an order of magnitude

(Rochette et al., 2008c). Once the crop N needs have been met, the

N2O emission rates appear to increase nonlinearly (Malhi et al., 2006;

McSwiney and Robertson, 2005; van Groenigen et al., 2010).

Anhydrous ammonia, urea, and urea ammonium nitrate (UAN) are the

most common forms of N fertilizer used in North America (Ribaudo et al.,

2011). A review of studies from around the world suggested that there is no



Nitrogenous Gas Emissions from Soils and Greenhouse Gas Effects



51



significant difference in N2O emissions from soils receiving these forms of N

fertilizers (Stehfest and Bouwman, 2006). However, side-by-side experiments in the United States and Canada suggest that the type of N fertilizer

applied can sometimes have a significant effect on the magnitude of N gas

emissions. In maize cropping systems, N2O emissions were higher from soils

receiving anhydrous ammonia than from those receiving UAN or urea

(Thornton et al., 1996; Gagnon et al., 2011; Fujinuma et al., 2011). The

difference in emissions between the two forms of N fertilizer may be due

to the placement/location of the fertilizer: anhydrous ammonia is injected

into the soil, whereas urea is broadcast. In a semiarid wheat cropping systems, however, no significant difference in N2O emissions was found in a

comparison of soils receiving either urea or anhydrous ammonia (Burton

et al., 2008), perhaps reflecting the lower N application rate in wheat systems

under dry climate.

Studies by Venterea et al. (2005, 2010) showed that tillage effects on

N2O emissions may depend on the form of N applied. They observed no

difference in N2O emissions between conventionally tilled and no-till systems when using UAN, but higher N2O emissions from soil under no-till

than from conventionally tilled soil when using broadcast urea. Emissions

following anhydrous ammonia were higher in conventionally tilled than

in no-till systems.

4.1.2 Synthetic FertilizersdTime of Application

Proper timing of N application is probably the biggest challenge in managing N fertilizer: matching adequate levels of soil mineral N with the uptake

of the added N by plants, without limiting their growth. Higher N2O emissions (as well as N leaching losses) will occur when soil mineral N exceeds

plant uptake. This synchrony is notoriously difficult in crops like maize

where the crop demand for N is very low early in the growing season,

then increasing rapidly during vegetative growth (i.e., V6-V7 stage), before

dropping sharply as the crop matures. Most N fertilizers are applied early,

either in fall or in spring, when there is no crop present or when the crop

is small and N demand is low, (e.g., about one-third of the maize in the

United States is fertilized in the fall (Paustian et al., 2004; Ribaudo et al.,

2011)), which can lead to large losses of N.

Improved synchrony may be achieved with side-dress fertilizer applications, which are applied later in the growing season and “enhancedefficiency fertilizers” (i.e., controlled-release N fertilizer; nitrification and

urease inhibitors), and a combination of these two approaches.



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4.1.3 Synthetic FertilizersdEffects on Microorganisms

Long-term N fertilization can increase or decrease microbial abundance and

population size of N2O producers and consumers. Changes in abundance

are primarily exerted by increasing the amount of plant-derived C inputs,

by short-term toxicity of concentrated N or through changes in soil pH,

which strongly interacts with N fertilizer to determine microbial abundance.

In soils of neutral to high pH, N fertilizer has a positive effect on microbial

abundance, while in soils with pH below 5, N application decreases microbial abundance (Geisseler and Scow, 2014). Since pH is also a strong determinant of community composition (Fierer and Jackson, 2006; Lauber et al.,

2009), the acidifying effect of N fertilization may influence N2O emissions

by changing the relative abundance of different N2O producers and

consumers.

Bacterial (AOB) and archaeal (AOA) ammonia oxidizers are thought to

have different pH optima (Nicol et al., 2008) with AOA dominating at low

pH. In contrast, AOB abundance is positively correlated with increasing pH

and these organisms may be absent in low pH environments (Prosser and

Nicol, 2012). Some evidence suggests that ammonia sensitivity is quite varied within both AOB and AOA and may govern activity (Jia and Conrad,

2009; Di et al., 2010; Yao et al., 2013). However, the short- versus longterm impacts of N fertilization on ammonia oxidizer community composition and activity are not well understood. Further evidence is needed to

unequivocally demonstrate the existence of different ecological niches for

the two groups (Taylor et al., 2012; Prosser and Nicol, 2012).

Type of N fertilizer can also affect microbial abundance and community

composition. Application of urea and ammonium salts leads to very high

concentrations of N (NH3 and NHỵ

4 ) toxic to microorganisms (Omar

and Ismail, 1999), but this appears to be short lived (Geisseler and Scow,

2014). Anhydrous ammonia fertilizer on the other hand appears to have

long-term toxic effects compared to urea on soil microorganisms and may

result in decreased microbial biomass, including nitrifiers (Biederbeck

et al., 1996).



4.2 Crop Residues and Organic Amendments

Incorporation of crop residues affects the emission of N gases from soil because

the N content of the residues controls mineralization/immobilization processes, thereby affecting the availability of N for nitrification/denitrification

processes. The magnitude of the emissions depends on the quantity,

quality, and timing of incorporation (Aulakh et al., 2001; Baggs et al.,



Nitrogenous Gas Emissions from Soils and Greenhouse Gas Effects



53



2003; Garcia-Ruiz and Baggs, 2007; Millar et al., 2004). Residues with relatively narrow C:N ratio have been shown to induce higher N2O emissions

than those in soils amended with residues having wide C:N (Toma and

Hatano, 2007; Gomes et al., 2009). These residue quality effects may last

for more than a single season; Liu et al. (2011) observed that incorporation

of wheat straw increased N2O and NO emissions in the following maize season, whereas incorporation of maize residues had no influence on emissions.

The effects of residues on substrate availability have been observed in

studies examining the microbial response to residue removal. For example,

Nemeth et al. (2014) found that removing crop residues increased N2O

emissions, which were correlated with lower copy number and transcription

of nosZ, suggesting that the increased emissions were the result of incomplete reduction of N2O to N2, rather than to an overall increase in the

amount of N denitrified. Furthermore, they showed that emissions in noresidue versus residue treatments were related to edaphic factors of C and

N availability during de novo denitrification throughout a spring thaw

event. Pelster et al. (2013) observed that return of crop residues decreased

N2O emissions by immobilizing N during freeze-thaw under controlled

conditions. Hence there are several lines of evidence to suggest that the

balance of N availability for denitrification with the timing of nosZ transcription is a critical factor in determining the N2:N2O ratio produced and that

this balance is coupled to residue decomposition.

Application of manure to soil usually increases N2O emissions through

stimulation of nitrification and denitrification processes (e.g., Chang et al.,

1998; Rochette et al., 2008a). Manure may contain large amounts of mineral and labile organic forms of N as well as large amounts of soluble organic

C, an important driver of denitrification. Response to manure addition may

therefore depend on the levels of available mineral N and soluble organic C

in the soil and in the manure.

Chantigny et al. (2010) observed that manure application led to lower

N2O emissions in a clay soil, but higher emissions in a loam soil, when

compared with N fertilizer application on the same soils. They attributed

the higher emissions in the loam soil to the availability of organic C in

the manure, which provided the substrate for denitrifying bacteria. Likewise

the level of available mineral N may be important in determining the

response to manure addition. Jarecki et al. (2009) observed that adding swine

manure increased cumulative N2O emissions but N availability was the

primary controlling mechanism because N2O emissions were not related

to %WFPS.



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Ed Gregorich et al.



Denitrification can also be induced as a result of increased oxygen consumption due to intense decomposition of manure shortly after its application (Smith et al., 2003). Thus relatively large emissions often occur within a

short time after application of manure (Whalen, 2000); for example,

Rochette et al. (2008a) reported that up to 60e90% of seasonal (i.e.,

MayeNovember) N2O emissions occurred within 40 days of manure application. But the manure effects on nitrification/denitrification processes can

be long lived. Ellert and Janzen (2008) detected an increase in N2O emissions over a 3-year period following a single manure application.

Application of biochar affects N gas emissions because it changes soil

physical and chemical properties such as soil structure, water holding

capacity, organic matter content, nutrient concentration, and pH

(Lehmann and Joseph, 2009). It can affect microbial N cycling (and thus

N gas production and emissions) through its effect on soil N concentration,

mineral N absorption, and competition with plants for N uptake (Clough

et al., 2013).



4.3 Tillage and Residue Management

The effect of tillage on N gas emission can be seen as a cascade of effects:

tillage affects soil structure and aeration, soil temperature, and water content.

All of these affect microbial activity, which in turn affects the rate of decomposition and N mineralization. However, the effects of tillage management

practices on N gas emission is not consistent; some studies have shown

decreased N2O emissions when no-till or reduced tillage is implemented

(e.g., Gregorich et al., 2007; Mosier et al., 2006), others reported higher

emissions (e.g., Ball et al., 1999; Burford et al., 1981), while still others

reported no difference (e.g., Lemke et al., 1999; Rochette et al., 2008b).

Some research has shown that the effects of tillage on N2O emissions depend

on the type of tillage management implemented and on placement of N fertilizer (Drury et al., 2006; Venterea et al., 2005).

From a review of literature Six et al. (2004) concluded that soil N2O

emissions usually increase during the first 10 years after tilled soils have

been converted to no-till; but this effect appeared to decrease over time

and this reduction in N2O emissions was apparent only in humid climates.

van Kessel et al. (2013), in a meta-analysis of field studies (239 comparisons),

showed that averaged across all comparisons, no-till/reduced till did not alter

N2O emissions compared with conventional tillage. However, no-till/

reduced till significantly reduced N2O emissions in experiments longer

than 10 years, especially in dry climates. The effects of tillage on soil N2O



Nitrogenous Gas Emissions from Soils and Greenhouse Gas Effects



55



emissions may be regulated to a large extent by climate. For example, in humid regions, like those in eastern Canada, soil N2O emissions under no-till

are more likely to be greater than those from conventional tillage, whereas in

semiarid regions, like those in western Canada, the converse may be true

(Helgason et al., 2005; Lemke and Janzen, 2007).

The gradually diminishing difference in N2O emission between tilled

and no-till systems may indicate the development of structure (Six et al.,

2004); the formation of macropores and water-stable aggregates over

time improves soil structure/aeration and decreases the formation of

anaerobic microsites where N2O is formed. Furthermore, the increase in

microbial biomass observed under long-term NT (Helgason et al., 2009)

may lead to an enhanced ability to immobilize N fertilizers at the time

of application, resulting in a more gradual release throughout the growing

season that is better synchronized with crop demand, particularly in dry

climates. Smith et al. (2010) showed that tillage management affects the

diversity of nitrifiers and denitrifiers and that these changes were apparent

during peak N2O emission events. However, a direct link between activity

of these communities (i.e., gene transcription) and N2O emissions was not

assessed.

van Kessel et al. (2013) found no significant correlation between soil

texture and the effect of no-till/reduced till on N2O emissions. This is

consistent with the findings of Skiba and Ball (2002) who concluded that

the local influences of soil structure/soil water status interactions and soil

mineral N content tend to override gross textural and drainage effects. These

findings contrast with those of Rochette et al. (2008a), who, in summarizing

25 field studies, observed that no-till generally increased N2O emissions in

poorly aerated soils but had no effect in soils with good and medium aeration. They concluded that the impact of no-till on N2O emissions is small in

well-aerated soils but most often positive in soils where aeration is reduced

by conditions or properties restricting drainage, such as is likely in no-till

soils.

Microbial abundance is tightly linked to C availability and thus, to the

quantity and quality of crop residues returned to the soil (Kallenbach and

Grandy, 2011). Both residue type and placement affect microbial community composition (Nicolardot et al., 2007) and active decomposer communities have been shown to differ between surface-placed and incorporated

residues (Helgason et al., 2014). Changes in residue decomposition under

NT and CT can alter substrate availability (C and N) for nitrification and

dentrification, affecting N2O emissions.



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4.4 Cropping Systems with Legumes

Symbiotic N fixation by legume crops contributes large amounts of N to

agricultural soils, and the flow of this N through the plantesoil system

can stimulate N2O production from several processes. Cropping systems

with legumes often produce lower annual N2O emission than fertilized

annual crops. The N2O emissions from legume crops are derived mainly

from decomposition of the above- and below-ground residues, and losses

from the biological N fixation process per se are likely negligible (Rochette

and Janzen, 2005).

Annual N2O-N emissions in alfalfa cropping systems have been shown

to be higher than those in soybean systems (Gregorich et al., 2005; Rochette

et al., 2004). This may be the result of frequent cutting and harvesting of

aboveground plant material on sources of N2O in soil (Rochette et al.,

2004), the litterfall of during the alfalfa growing season (Tomm et al.,

1995); as well, biological N fixation may be higher in perennial alfalfa

than in soybean. This is consistent with the observation that the dieback

of alfalfa nodules occurs following harvest (Vance et al., 1979), which could

contribute to N release from the root systems. Wagner-Riddle et al. (1997)

measured high N2O emissions in the spring following plow down of an

alfalfa crop the previous autumn. This suggests that total N2O emissions

in alfalfa cropping systems may be greater than those only measured during

the growing season. This highlights the importance of processes that

contribute to the production of N2O following harvest and plow down

of N-rich crop residues.



4.5 Rotation

Rotation effects on N2O emissions relate mainly to the amount of fertilizer

N and the quantity and quality of crop residues returned to the soil. MeyerAurich et al. (2006) compared estimated N2O emissions in six maize-based

rotations under moldboard and chisel plow systems in a 20-year field experiment in eastern Ontario. Differences in emissions between tillage systems

were small in comparison to rotation effects; crop rotations with legumes

had substantially lower N2O emissions than continuous maize, reflecting

the higher inputs of fertilizer N in the monoculture system.

Maizeesoybean rotations often have lower cumulative N2O emissions

compared to continuous maize production, because of lower seasonal

N2O emission from soybean than from maize (Rochette et al., 2004; Mosier

et al., 2006; Hernandez-Ramierz et al., 2009). Carryover of N2O emissions



Nitrogenous Gas Emissions from Soils and Greenhouse Gas Effects



57



from corn to soybean years within the rotation seems minimal (Grandy et al.,

2006; Mosier et al., 2006; Parkin and Kaspar, 2006; Hernandez-Ramierz

et al., 2009).



4.6 Irrigation and Drainage

Irrigation typically is assumed to increase direct N2O emissions systems

because water-filled pore space is expected to be greater than that in rainfed systems. Studies to verify such assumptions, however, are scant because

irrigation tends to confound comparisons. Since irrigated systems tend to be

more productive, greater amounts of reactive N are circulating. Furthermore, irrigation water tends to be applied when transpiration rates are

high, so periods when soils are moistened by irrigation often do not coincide

with periods of high fluxes (Ellert and Janzen, 2008). Most studies do not

directly compare the same crop with and without irrigation, but in a multiyear study of maize in Colorado, USA, soil N2O emissions were smaller in a

year with greater inputs of rainfall plus irrigation even though the latter

culminated in greater proportions of water-filled soil pores (Halvorson

and Del Grosso, 2013). In a recent review of soil N2O emissions from irrigated and nonirrigated systems, Trost et al. (2013) concluded that the primary factor responsible for increased emissions from irrigated systems was

usually increased amounts of available soil N, rather than increased waterfilled soil pores. Since soil N2O is a transitory intermediary, the balance

between production and consumption is not expected to follow a

straight-forward relationship with water-filled soil pores, as influenced by

irrigation. Between 20% and 60% water-filled pore space, N2O leakage during nitrification may predominate; from 60% to 75% water-filled pore space,

N2O from denitrification may predominate; but at greater than 75%, consumption by denitrification may even decrease soil N2O fluxes. After

accounting for the significantly increased yields with irrigation observed in

many studies (e.g., Maharjan et al., 2014), N losses per unit of yield did

not differ and in some cases decreased with irrigation.



5. AGRONOMIC ASSESSMENT OF N GAS EMISSIONS

AND BROADER ENVIRONMENTAL CONTEXT OF

N FERTILIZERS

5.1 Yield-Scaled Emissions

To satisfy the increased global demand for agricultural products, it will

be necessary to strike a judicious balance between an agronomic perspective



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Ed Gregorich et al.



of maintaining adequate N inputs and an environmental perspective of

minimizing N losses from soil to the atmosphere. This concept is embodied

in the approach where N2O emissions are expressed as a function of crop

productivity measurements like yield or N uptake (van Groenigen et al.,

2010; Venterea et al., 2011). Thus N2O emissions are thought of as a

“cost of production,” and the aim is to maximize returns per unit “cost.”

Scaling or normalizing the emissions to the quantity of crop produced helps

put the emissions “cost” into perspective and allows more holistic comparison of management systems.

The yield-scaled N2O emission metric is helpful to evaluate the effects of

soil degradation on N2O emissions because it captures both the negative

environmental and agronomic aspects of soil compaction in a single parameter. Gregorich et al. (2014) observed that yield-scaled N2O emissions were

particularly sensitive to compaction because compaction both increased the

area-scaled N2O emissions (metric numerator) and decreased the yield

(metric denominator) providing a double effect on the yield-scaled N2O

emission metric. Consequently, the ratios in yield-scaled N2O emissions

between compacted and not compacted soils were higher than the ratios

in area-scaled N2O emissions in the same year. In contrast, increasing N fertilizer increases both area-scaled N2O emissions and yields. As a result, the

ratio of yield-scaled N2O emissions between high and low N fertilizer rates

were less than the ratio of area-scaled N2O emissions for high and low N

fertilizer rates in the same year.



5.2 Fertilizer Use and Efficiency in Developing versus

Developed Countries

Since 1960, N fertilizer use has steadily increased in all countries but the rate of

increase has slowed in developed countries since about 1980, while in developing countries it has increased during this period (Figure 7; FAO, 2011). Recovery of N fertilizers by grain (maize, rice, and wheat) crops in field research

plots can be as high as 85% but is generally 30e40% (Krupnik et al., 2004).

Across regions of the world the average recovery for grain crops is lowest in

Africa (26%) and this low efficiency use can be attributed to other limiting

factors such as water and phosphorus deficiency (Krupnik et al., 2004).

Across agricultural production systems the current fertilizer-N use efficiency

is estimated to be about 50% (Smil, 1999; Ladha et al., 2005).

These losses of more than 50% of the applied fertilizer are dissipated

across the wider environment by leaching of soluble N and emission of N

gases to the atmosphere. The losses also represent significant economic



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