**Greenhouse Gases Emission and Measurement**

**1** 

*New Zealand* 

**Emissions of Nitrous Oxide (N2O)** 

**Affecting N2O and N2 Ratios** 

M.J. Khan, M.N. Babar and S. Zaman *Ballance Agri-Nutrients Limited, Tauranga,* 

M. Zaman, M.L. Nguyen, M. Šimek, S. Nawaz,

**and Di-Nitrogen (N2) from the Agricultural Landscapes, Sources, Sinks, and Factors** 

Nitrous oxide (N2O) is one of the key greenhouse and ozone (O3) depleting gas, constituting 7% of the anthropogenic greenhouse effect. On a molecular basis, N2O has 310 and 16 times higher global warming potential than that of CO2 and CH4 respectively over a 100-year period. To develop mitigation tools for N2O emissions, it is imperative to understand the processes of nitrogen (N) transformation and N2O and di-nitrogen (N2) production in soils as influenced by different land uses, management and environmental conditions. The aim of our chapter is to examine the current information and understanding of the sources of N2O and N2 production and the factors affecting N2O:N2 ratio from the agricultural landscapes. Nitrous oxide concentration has increased by 20% from 270 ppbv since 1750 to a current level of 322 ppbv and continues to increase currently by 0.3% per year. Intensification of agricultural and human activities, such as the increased use of synthetic fertilizer (103 M ton of N worldwide in 2010), increasing human population and changes in their diet, inefficient use of irrigation water, increased crop production, deposition of animal excreta (urine + dung) from grazing animals, excessive application rates of farm effluents and animal manures to croplands and pastures, and management practices that enhance soil organic N mineralization and C decomposition including cultivation, residues removal or burning, and following no crop rotation are to be blamed for the increased N2O emissions of 17.7 T g of N per year to the atmosphere. This book chapter focuses on the following sub-sections including nitrogen transformations, processes of N2O and N2 production across the agricultural landscape, challenges in N2O measurements and estimates across the agricultural landscape, factors affecting N2O and N2 emissions and possible mitigating

Nitrogen is an essential nutrient controlling the diversity, dynamics, and functioning of many terrestrial, freshwater and marine ecosystems. Agricultural ecosystems rely on N

**1. Introduction** 

options, conclusions and references.

**2. Nitrogen transformations** 

## **Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios**

M. Zaman, M.L. Nguyen, M. Šimek, S. Nawaz, M.J. Khan, M.N. Babar and S. Zaman *Ballance Agri-Nutrients Limited, Tauranga, New Zealand* 

#### **1. Introduction**

Nitrous oxide (N2O) is one of the key greenhouse and ozone (O3) depleting gas, constituting 7% of the anthropogenic greenhouse effect. On a molecular basis, N2O has 310 and 16 times higher global warming potential than that of CO2 and CH4 respectively over a 100-year period. To develop mitigation tools for N2O emissions, it is imperative to understand the processes of nitrogen (N) transformation and N2O and di-nitrogen (N2) production in soils as influenced by different land uses, management and environmental conditions. The aim of our chapter is to examine the current information and understanding of the sources of N2O and N2 production and the factors affecting N2O:N2 ratio from the agricultural landscapes. Nitrous oxide concentration has increased by 20% from 270 ppbv since 1750 to a current level of 322 ppbv and continues to increase currently by 0.3% per year. Intensification of agricultural and human activities, such as the increased use of synthetic fertilizer (103 M ton of N worldwide in 2010), increasing human population and changes in their diet, inefficient use of irrigation water, increased crop production, deposition of animal excreta (urine + dung) from grazing animals, excessive application rates of farm effluents and animal manures to croplands and pastures, and management practices that enhance soil organic N mineralization and C decomposition including cultivation, residues removal or burning, and following no crop rotation are to be blamed for the increased N2O emissions of 17.7 T g of N per year to the atmosphere. This book chapter focuses on the following sub-sections including nitrogen transformations, processes of N2O and N2 production across the agricultural landscape, challenges in N2O measurements and estimates across the agricultural landscape, factors affecting N2O and N2 emissions and possible mitigating options, conclusions and references.

#### **2. Nitrogen transformations**

Nitrogen is an essential nutrient controlling the diversity, dynamics, and functioning of many terrestrial, freshwater and marine ecosystems. Agricultural ecosystems rely on N

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

as shown in Eq. 2.1.1

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 5

immobilization process is the opposite of mineralization, where mineral N (NH4+ and NO3-

and even organic N (amino acids) are consumed by a diverse group of microorganisms to synthesize their protein and grow in number. These diverse groups of microorganisms are mostly heterotrophic (consume organic C) bacteria, fungi, and actinomycete, which produce a wide variety of extracellular and intra-cellular enzymes (e.g. protease, deaminase and urease) in soil. They then carry out the hydrolysis of high molecular weight organic compounds like protein into low molecular weight organic compounds such as amino acids,

(2.1.1) The low molecular weight organic compounds such as amino acids, amines and amides produced after proteolytic decomposition or after the application of organic residues or wastes are then subjected to microbial and enzymatic decomposition such as deamination which is carried out either by extracellular deaminase (Ladd & Jackson, 1982; Zaman et al., 1999 a, b) or by direct assimilation within the microbial cell (Barak et al., 1990; Barraclough, 1997). A large number of heterotrophic microorganisms are capable of carrying out the deamination of amino acids, both within and outside the microbial cell. Deaminases hydrolyse the NH2-N attached to the -C of an amino acid to NH3 and CO2. The amino acids are deaminated at different rates through four different reactions, depending on their side chains, as shown below in Eq. 2.1.2. Some amino acids are reported to be readily

mineralized, while others take longer to mineralize (Alef & Kleiner, 1986).

)

inputs from a variety of sources including synthetic chemical fertilizers, predominantly urea which accounts for more than 50% of the total world N consumption, organic wastes (farm dairy effluent, animal excreta, plant residues and sewage sludge) and atmosphere (biological fixation of atmospheric N through symbiotic and non-symbiotic microorganisms) to sustain productivity. A detailed description of N cycling in agricultural ecosystems is beyond the scope of this chapter and for details on N transformations, N dynamics, sources of N inputs, and losses, the readers are referred to research papers, articles and review written by these authors (Ledgard et al., 1999; Saggar et al., 2004b, 2005, 2009, 2011; deKlein & Eckard 2008; Ledgard & Luo 2008; Luo et al., 2010); however a brief description of the various microbial and enzymatic processes involved in N cycling is given below.

#### **2.1 A brief biochemistry of N mineralization**

Nitrogen transformations within soil-plant-water and atmospheric systems refer to N cycling. As will be discussed in section 3, N cycling provides precursors like ammonium (NH4+) and nitrate (NO3- ) for the production of N2O and N2 in soil. A simple schematic diagram of the N inputs, losses and transformation processes is presented in Fig. 1 The key N transformation processes within soil, plant and atmospheric systems include mineralization (gross and net), immobilization, nitrification (gross and net), denitrification, ammonia (NH3) volatilization, NH4+ fixation and NO3 leaching. The first four processes (i.e. mineralization, immobilization, nitrification and denitrification) are of microbial and enzymatic origin (biotic), while the last three (i.e. NH3 volatilization, NH4 + fixation and NO3 leaching) involve only chemical and physical processes (abiotic). Nitrogen mineralization is a sequence of microbial and enzymatic activities which involves the conversion of organic N (eg. protein, amino acids, amines, amides, urea, chitin and amino sugars) into an inorganic form of N (mainly NH4+), which then serves as a substrate for a diverse group of microorganisms and for nitrification (Zaman et al., 1999 a, b; 2004; Zaman & Change, 2004). The

Fig. 1. N inputs, losses and transformation processes across the agricultural landscape (Zaman et al., 2008b).

inputs from a variety of sources including synthetic chemical fertilizers, predominantly urea which accounts for more than 50% of the total world N consumption, organic wastes (farm dairy effluent, animal excreta, plant residues and sewage sludge) and atmosphere (biological fixation of atmospheric N through symbiotic and non-symbiotic microorganisms) to sustain productivity. A detailed description of N cycling in agricultural ecosystems is beyond the scope of this chapter and for details on N transformations, N dynamics, sources of N inputs, and losses, the readers are referred to research papers, articles and review written by these authors (Ledgard et al., 1999; Saggar et al., 2004b, 2005, 2009, 2011; deKlein & Eckard 2008; Ledgard & Luo 2008; Luo et al., 2010); however a brief description of the various

Nitrogen transformations within soil-plant-water and atmospheric systems refer to N cycling. As will be discussed in section 3, N cycling provides precursors like ammonium

diagram of the N inputs, losses and transformation processes is presented in Fig. 1 The key N transformation processes within soil, plant and atmospheric systems include mineralization (gross and net), immobilization, nitrification (gross and net), denitrification, ammonia (NH3) volatilization, NH4+ fixation and NO3- leaching. The first four processes (i.e. mineralization, immobilization, nitrification and denitrification) are of microbial and

leaching) involve only chemical and physical processes (abiotic). Nitrogen mineralization is a sequence of microbial and enzymatic activities which involves the conversion of organic N (eg. protein, amino acids, amines, amides, urea, chitin and amino sugars) into an inorganic form of N (mainly NH4+), which then serves as a substrate for a diverse group of microorganisms and for nitrification (Zaman et al., 1999 a, b; 2004; Zaman & Change, 2004). The

> Org.N NH 4 +

N O - NO N O 2

N 2 NO 2

**Denitrification**

**Seepage**

NO 3 -

**Pasture soil Wetland Stream**

 NO 2 -

**Org.N, NH , No 4 3 + -**

**Surface runoff**

2

**Nitrification**

Fig. 1. N inputs, losses and transformation processes across the agricultural landscape

) for the production of N2O and N2 in soil. A simple schematic

+ fixation and NO3-

Org.N NH 4 +

 N2 NO 2 - NO N O 2

**Denitrification**

NO 3 -

 NO 2 -

 2 N O

**Nitrification**

microbial and enzymatic processes involved in N cycling is given below.

enzymatic origin (biotic), while the last three (i.e. NH3 volatilization, NH4

Rainfall N fixation

N O 2

**Nitrification**

Animal Effluent 2

N Fertilizer **NH4 +**

Org.N NH 4 +

> NO 2 -

NO 3 -

**Org.N, NH , NO 4 3 + -**

NO 2


**Denitrification**

**Leaching**

(Zaman et al., 2008b).

N Fertilizer

> **NO3 -**

irrigation **Nitrogen addition**

**2.1 A brief biochemistry of N mineralization** 

(NH4+) and nitrate (NO3-

immobilization process is the opposite of mineralization, where mineral N (NH4 + and NO3 - ) and even organic N (amino acids) are consumed by a diverse group of microorganisms to synthesize their protein and grow in number. These diverse groups of microorganisms are mostly heterotrophic (consume organic C) bacteria, fungi, and actinomycete, which produce a wide variety of extracellular and intra-cellular enzymes (e.g. protease, deaminase and urease) in soil. They then carry out the hydrolysis of high molecular weight organic compounds like protein into low molecular weight organic compounds such as amino acids, as shown in Eq. 2.1.1

The low molecular weight organic compounds such as amino acids, amines and amides produced after proteolytic decomposition or after the application of organic residues or wastes are then subjected to microbial and enzymatic decomposition such as deamination which is carried out either by extracellular deaminase (Ladd & Jackson, 1982; Zaman et al., 1999 a, b) or by direct assimilation within the microbial cell (Barak et al., 1990; Barraclough, 1997). A large number of heterotrophic microorganisms are capable of carrying out the deamination of amino acids, both within and outside the microbial cell. Deaminases hydrolyse the NH2-N attached to the -C of an amino acid to NH3 and CO2. The amino acids are deaminated at different rates through four different reactions, depending on their side chains, as shown below in Eq. 2.1.2. Some amino acids are reported to be readily mineralized, while others take longer to mineralize (Alef & Kleiner, 1986).

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 7

The active site of urease contains two-nickel (II) atoms, which are linked by a carbamate bridge. Two imidazole N atoms are bound to each Ni atom; a carboxylate group and a water molecule fill the remaining coordination site of the metal ion. The ability to hydrolyze urea is found to vary from 17 to 70% for soil bacteria and from 78 to 98% for soil fungi (Lloyd & Sheaffe, 1973; Roberge & Knowles, 1967). Although soil urease is considered to be of microbial origin (Skujins, 1976), there is evidence to suggest that some urease activity may be derived from plants (Frankenberger & Tabatabai, 1982). However, there is no direct

evidence for the production of urease by plant roots (Estermann & McLaren, 1961).

**3. Processes of N2O and N2 production across the agricultural landscape** 

Gaseous N emissions from the agricultural landscape (arable, pasture and wetland soils) occur as NH3, nitric oxide (NO also called nitrogen oxide), nitrogen dioxide (NO2), N2O and N2. Quantifying N2O emission is of particular interest to those countries, which are signatories to the Kyoto Protocol, since it is one of the key greenhouse gases constituting 7% of the anthropogenic greenhouse effect. On a molecular basis, N2O has 310 and 16 times higher global warming potential than that of CO2 and methane (CH4) respectively over a 100-year period (IPCC, 2007). The global atmospheric concentration of N2O has increased from 270±7 in the pre-industrial-period (1750) to a current level of 322 ppbv representing a 20% increase. Over the last two decades a nearly linear increase of 0.26% in the concentration of N2O has been measured (Saggar et al., 2009). Moreover, due to its relative stability, (150 years) after emission from the soil surface and transport through the troposphere, N2O acts as a source of NO in the stratosphere, and thus indirectly accelerates depletion of ozone (O3), a substance that protects the biosphere from harmful ultraviolet (UV) radiation (Crutzen, 1981; Duxbury, 1994). The total estimated emissions of N2O are about 17.7 Tg N per year, but there are large uncertainty ranges in each of the individual sources. About 70% of N2O emissions come from the bacterial breakdown of N in soils and in the oceans. Globally, soils in areas of natural vegetation, especially in the tropics, and the oceans account for N2O emissions of about 6.6 and 3.8 Tg N per year respectively; while human activities account for the remaining 30% of N2O emissions, or about 6.7 Tg N per year (Denman, 2007). Factors blamed for the increased N2O emissions of 17.7 T g of N per year to the atmosphere include; a rapid increase in human population (according to the latest United Nations population estimates, 77 million more people each year are being added to the current world population of 6.98 billion), intensification of agricultural and human activities, such as the increased use of synthetic fertilizer (103 million ton of N worldwide in 2010) (IFA 2011), inefficient use of irrigation water and N fertilizers (both synthetic and organic), increased grassland areas for livestock which cover 117 million km2 of vegetated lands that provide forage for over 1800 million livestock units and wildlife (World Resources Institute 2000). The other factors include increased animal stocking rates (>3 cows per ha) and intensive gazing, which results in deposition of huge amounts of N via animal excreta (urine + dung), farm management practices that enhance soil organic N mineralization and decomposition of organic C (deep cultivation, crop residues removal or burning, and following no crop rotation) and the increased consumption of dairy products worldwide especially in fast growing economies like China and India (Robertson et al., 1989; Duxbury et al., 1993; Šimek & Cooper, 2001; Rochester, 2003; Denman et al*.,* 2007; IPCC, 2010; Zaman & Blennerhassett., 2010; Zaman & Nguyen., 2010). Nitrous oxide can also be produced during nitric acid production and fossil fuel combustion, but the amount of N2O

Whether an amino acid is used for an energy source by microorganisms or as a building block for protein synthesis depends on the available N and soluble organic C at the microsite where the microbial reaction occurs (Mengel, 1996). After deamination has occurred within the cell, the removal of NH4+ is carried out by enzymes such as glutamate dehydrogenase and coenzyme nicotine adenine dinucleotide (NADH). Ammonium produced by deamination is always associated with the production of new microbial biomass, and the extent of NH4+ immobilization or accumulation in the soil depends on the micro-organism's C:N ratio (Mengel, 1996; Paul & Clark, 1996) and the available soil mineral N and organic C. The turnout of microbial biomass is reported to be fast and the new microbial biomass die after reaching a certain limit, thus serving as a substrate for enzymes and other groups of microorganisms. This turnover of microorganisms releases the NH4+ again. Thus the dead biomass, which is prone to biological decomposition (Jenkinson & Ladd, 1981), serves as the main source of NH4+ production in soil (Azam et al., 1986). The non-proteinaceous cell wall constituents of bacteria and fungi, such as amino sugar and chitin, are first depolymerised by chitinase to glucose amines. These are then attacked by kinases, and this process finally releases the NH4+ in soil as shown in Eq. 2.1.3.

$$\text{(21.3)}$$

Similarly urea (CO (NH2)2, from (i) urine deposition of grazing animals, (ii) the application of urea fertilizer or (iii) from production of hydrolytic decomposition of proteinaceous materials in soil, undergoes fast hydrolysis (Zaman et al., 2008a & 2009) and the hydrolysis is usually completed within 1 to 2 days by urease enzymes. These ubiquitous enzymes are found in soils, many plants and plant litters (Freney & Black, 1988) and in most species of bacteria, yeast and fungi (Sumner, 1953). Urease catalyzes the hydrolysis of urea to NH4+ (Eq. 2.1.4) and carbamate ions, which result in the production of carbon dioxide (CO2) and NH4+.

$$\begin{array}{ccccc} \underset{\text{H}\_{2}\text{N}^{+}}{\underset{\text{NH}\_{2}}{\rightleftharpoons}} & + & \quad \text{2H}\_{2}\text{O} & \xrightarrow{\text{Soll Unase}} & \quad \text{(NH}\_{3}\text{CO}\_{3} \\\\ & & & & \text{2H}^{+} \\\\ & & & & \text{2NH}\_{4}^{+} + \text{CO}\_{2} & + & \text{H}\_{2}\text{O} \end{array} \quad \text{(NH}\_{3}\text{O}\_{2}\text{O}\_{3})$$

(2.1.2) Whether an amino acid is used for an energy source by microorganisms or as a building block for protein synthesis depends on the available N and soluble organic C at the microsite where the microbial reaction occurs (Mengel, 1996). After deamination has occurred within the cell, the removal of NH4+ is carried out by enzymes such as glutamate dehydrogenase and coenzyme nicotine adenine dinucleotide (NADH). Ammonium produced by deamination is always associated with the production of new microbial biomass, and the extent of NH4+ immobilization or accumulation in the soil depends on the micro-organism's C:N ratio (Mengel, 1996; Paul & Clark, 1996) and the available soil mineral N and organic C. The turnout of microbial biomass is reported to be fast and the new microbial biomass die after reaching a certain limit, thus serving as a substrate for enzymes and other groups of microorganisms. This turnover of microorganisms releases the NH4+ again. Thus the dead biomass, which is prone to biological decomposition (Jenkinson & Ladd, 1981), serves as the main source of NH4+ production in soil (Azam et al., 1986). The non-proteinaceous cell wall constituents of bacteria and fungi, such as amino sugar and chitin, are first depolymerised by chitinase to glucose amines. These are then attacked by

kinases, and this process finally releases the NH4+ in soil as shown in Eq. 2.1.3.

NH4+.

(2.1.3) Similarly urea (CO (NH2)2, from (i) urine deposition of grazing animals, (ii) the application of urea fertilizer or (iii) from production of hydrolytic decomposition of proteinaceous materials in soil, undergoes fast hydrolysis (Zaman et al., 2008a & 2009) and the hydrolysis is usually completed within 1 to 2 days by urease enzymes. These ubiquitous enzymes are found in soils, many plants and plant litters (Freney & Black, 1988) and in most species of bacteria, yeast and fungi (Sumner, 1953). Urease catalyzes the hydrolysis of urea to NH4+ (Eq. 2.1.4) and carbamate ions, which result in the production of carbon dioxide (CO2) and

(2.1.4)

The active site of urease contains two-nickel (II) atoms, which are linked by a carbamate bridge. Two imidazole N atoms are bound to each Ni atom; a carboxylate group and a water molecule fill the remaining coordination site of the metal ion. The ability to hydrolyze urea is found to vary from 17 to 70% for soil bacteria and from 78 to 98% for soil fungi (Lloyd & Sheaffe, 1973; Roberge & Knowles, 1967). Although soil urease is considered to be of microbial origin (Skujins, 1976), there is evidence to suggest that some urease activity may be derived from plants (Frankenberger & Tabatabai, 1982). However, there is no direct evidence for the production of urease by plant roots (Estermann & McLaren, 1961).

## **3. Processes of N2O and N2 production across the agricultural landscape**

Gaseous N emissions from the agricultural landscape (arable, pasture and wetland soils) occur as NH3, nitric oxide (NO also called nitrogen oxide), nitrogen dioxide (NO2), N2O and N2. Quantifying N2O emission is of particular interest to those countries, which are signatories to the Kyoto Protocol, since it is one of the key greenhouse gases constituting 7% of the anthropogenic greenhouse effect. On a molecular basis, N2O has 310 and 16 times higher global warming potential than that of CO2 and methane (CH4) respectively over a 100-year period (IPCC, 2007). The global atmospheric concentration of N2O has increased from 270±7 in the pre-industrial-period (1750) to a current level of 322 ppbv representing a 20% increase. Over the last two decades a nearly linear increase of 0.26% in the concentration of N2O has been measured (Saggar et al., 2009). Moreover, due to its relative stability, (150 years) after emission from the soil surface and transport through the troposphere, N2O acts as a source of NO in the stratosphere, and thus indirectly accelerates depletion of ozone (O3), a substance that protects the biosphere from harmful ultraviolet (UV) radiation (Crutzen, 1981; Duxbury, 1994). The total estimated emissions of N2O are about 17.7 Tg N per year, but there are large uncertainty ranges in each of the individual sources. About 70% of N2O emissions come from the bacterial breakdown of N in soils and in the oceans. Globally, soils in areas of natural vegetation, especially in the tropics, and the oceans account for N2O emissions of about 6.6 and 3.8 Tg N per year respectively; while human activities account for the remaining 30% of N2O emissions, or about 6.7 Tg N per year (Denman, 2007). Factors blamed for the increased N2O emissions of 17.7 T g of N per year to the atmosphere include; a rapid increase in human population (according to the latest United Nations population estimates, 77 million more people each year are being added to the current world population of 6.98 billion), intensification of agricultural and human activities, such as the increased use of synthetic fertilizer (103 million ton of N worldwide in 2010) (IFA 2011), inefficient use of irrigation water and N fertilizers (both synthetic and organic), increased grassland areas for livestock which cover 117 million km2 of vegetated lands that provide forage for over 1800 million livestock units and wildlife (World Resources Institute 2000). The other factors include increased animal stocking rates (>3 cows per ha) and intensive gazing, which results in deposition of huge amounts of N via animal excreta (urine + dung), farm management practices that enhance soil organic N mineralization and decomposition of organic C (deep cultivation, crop residues removal or burning, and following no crop rotation) and the increased consumption of dairy products worldwide especially in fast growing economies like China and India (Robertson et al., 1989; Duxbury et al., 1993; Šimek & Cooper, 2001; Rochester, 2003; Denman et al*.,* 2007; IPCC, 2010; Zaman & Blennerhassett., 2010; Zaman & Nguyen., 2010). Nitrous oxide can also be produced during nitric acid production and fossil fuel combustion, but the amount of N2O

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 9

(3.1.2)

presence of O2 and organic C, can also produce N2O from NO2- and typically occurs in acidic soils (Wood, 1990). However, high rates of heterotrophic nitrification relative to autotrophic nitrification have been measured in a riparian wetland soil with a pH close to 7, which was exposed to O2 (Matheson et al., 2003). Production of N2O via heterotrophic nitrification is poorly understood because autotrophic and heterotrophic nitrification can occur simultaneously in a given soil and it is difficult to separate the end products of these two processes without using 15N tracers (Matheson et al., 2003). Sufficient soil O2 levels

favorable soil temperature above 5oC (optimum 25 to 35oC), and soil pH above 5 (optimum 7 to 9) are among the known soil and environmental conditions which control the rate of autotrophic nitrification (Linn & Doran 1984; Grundmann et al., 1995; Whitehead, 1995; Zaman et al., 1999a; Šimek., 2000; Zaman & Chang, 2004; Zaman et al., 2007; Saggar et al.,

concentrations are considered the most critical factors affecting autotrophic nitrification (Zaman et al., 2007). Thus autotrophic nitrification is expected to be a dominant N transformation process in well-drained pastoral or arable systems, where soils are oxygenated (at or around field capacity or at 60% WFPS) and NH4+ is abundant [(e.g., excreta deposition after animal grazing, after the application of organic wastes, and NH4+ based synthetic fertilizer like urea, di-ammonium phosphate (DAP), ammonium sulphate, and liquid ammonia or as a result of increased mineralization of soil organic N compounds)] (Zaman et al., 1999a,b; Zaman & Chang, 2004; Zaman et al., 2007; 2008a; 2009; Zaman & Nguyen 2010). However, nitrification can also occur in waterlogged areas at a slower rate where wetland vegetation releases O2 from roots (Armstrong, 1964). At the sediment-plant root interface, nitrifying bacteria are supplied with O2 from plants and NH4+ from the surrounding sediment. There is evidence to suggest that autotrophic denitrification can proceed at a pH around 4, because soil aggregates protect bacterial cells against nitrous

N2O and N2 in a respiratory metabolism. During respiratory denitrification, denitrifiers couple reduction of N-oxides to oxidation of organic C under anaerobic conditions and

2009; Zaman et al., 2009; Zaman & Nguyen, 2010). Among these factors, NH4

+ to NO3- in the

+ concentrations, a

and NO2


+ and O2

Broken lines show the unconfirmed pathways of the biological reaction.

[(optimum at water filled pore space (WFPS) of 60%)], adequate NH4

Denitrification is predominantly a microbial process by which NO3-

acid toxicity (De Boer et al., 1991).

**3.2 Denitrification** 

Heterotrophic nitrification, the oxidation of reduced N compounds or NH4

from fossil fuel varies with the fuel type and technology. Fossil fuel combustion and industrial processes are responsible for N2O emissions of around 0.7 Tg N per year (Denman, 2007). Other important sources include human sewage and the burning of biomass and biofuels.

Across the agricultural landscape, several microbial processes can occur simultaneously to produce harmful N2O and non-greenhouse N2 in soils (pasture and arable) and sediments (drain, ditch, wetland and stream). These microbial processes are regulated by various soil, environmental and management factors, therefore making it difficult to control the rates of N2O and N2 production and their ratios (Paul & Beauchamp, 1989; Stevens et al., 1997; Zaman et al., 2007, 2008 b,c; Zaman & Nguyen, 2010). The aim of our review is to examine current information and understanding of the sources of N2O and N2 production and factors affecting N2O:N2 ratio in agricultural landscape, to enable management practices to be devised that minimize N losses as N2O emission to the atmosphere.

Soil microbial processes, which account for major N2O production include; nitrification (Inubushi et al., 1996), denitrification (Tiedje, 1988; Firestone & Davidson, 1989; Smith, 1990; Cavigelli & Robertson, 2001) and dissimilatory NO3 - reduction to NH4 + (DNRA) (Silver et al., 2001). These three microbial processes may occur in soils and sediments across the landscape depending on the physical (moisture contents or O2 level) and chemical conditions [N form (i.e. NH4 + and NO3-), pH and C contents] in their micro-sites. Details of each of these processes are given below:

#### **3.1 Nitrification**

Autotrophic nitrification, a strictly aerobic process, is carried by chemolitho-autotrophic bacteria which use O2 as a terminal electron acceptor. In the first step, NH4+ is oxidized to NO2- by ammonia oxidizing species of the genus *Nitrosomonas,* while in the second step, NO2- oxidation to NO3- is facilitated by *Nitrobacter* and *Nitrococcus* (Bremner & Blackmer, 1981; Watson et al. 1981) as shown in Eq. 3.1.1 Other genera including *Nitrosococus, Nitrosospira* and subgenera *Nitrosobolus, and Nitrosovibrio* also have the ability to autotrophically oxidize NH3 to NO2-:

In addition to NO2- production during the first stage of autotrophic nitrification, several intermediate and unstable compounds such as hydroxylamine (NH2OH) and nitroxyl (NOH) are also formed. Ammonia oxidizers consume relatively large amounts of molecular O2 during this first stage, causing anaerobic conditions in the microsites, which then lead to a reduction of NO2 - to N2O and N2 (Poth & Focht, 1985; Firestone & Davidson, 1989; Zart & Bock, 1998; Colliver & Stephenson, 2000) as shown in Eq. 3.1.2.

(3.1.2)

Broken lines show the unconfirmed pathways of the biological reaction.

Heterotrophic nitrification, the oxidation of reduced N compounds or NH4 + to NO3 - in the presence of O2 and organic C, can also produce N2O from NO2- and typically occurs in acidic soils (Wood, 1990). However, high rates of heterotrophic nitrification relative to autotrophic nitrification have been measured in a riparian wetland soil with a pH close to 7, which was exposed to O2 (Matheson et al., 2003). Production of N2O via heterotrophic nitrification is poorly understood because autotrophic and heterotrophic nitrification can occur simultaneously in a given soil and it is difficult to separate the end products of these two processes without using 15N tracers (Matheson et al., 2003). Sufficient soil O2 levels [(optimum at water filled pore space (WFPS) of 60%)], adequate NH4 + concentrations, a favorable soil temperature above 5oC (optimum 25 to 35oC), and soil pH above 5 (optimum 7 to 9) are among the known soil and environmental conditions which control the rate of autotrophic nitrification (Linn & Doran 1984; Grundmann et al., 1995; Whitehead, 1995; Zaman et al., 1999a; Šimek., 2000; Zaman & Chang, 2004; Zaman et al., 2007; Saggar et al., 2009; Zaman et al., 2009; Zaman & Nguyen, 2010). Among these factors, NH4 + and O2 concentrations are considered the most critical factors affecting autotrophic nitrification (Zaman et al., 2007). Thus autotrophic nitrification is expected to be a dominant N transformation process in well-drained pastoral or arable systems, where soils are oxygenated (at or around field capacity or at 60% WFPS) and NH4 + is abundant [(e.g., excreta deposition after animal grazing, after the application of organic wastes, and NH4+ based synthetic fertilizer like urea, di-ammonium phosphate (DAP), ammonium sulphate, and liquid ammonia or as a result of increased mineralization of soil organic N compounds)] (Zaman et al., 1999a,b; Zaman & Chang, 2004; Zaman et al., 2007; 2008a; 2009; Zaman & Nguyen 2010). However, nitrification can also occur in waterlogged areas at a slower rate where wetland vegetation releases O2 from roots (Armstrong, 1964). At the sediment-plant root interface, nitrifying bacteria are supplied with O2 from plants and NH4 + from the surrounding sediment. There is evidence to suggest that autotrophic denitrification can proceed at a pH around 4, because soil aggregates protect bacterial cells against nitrous acid toxicity (De Boer et al., 1991).

#### **3.2 Denitrification**

8 Greenhouse Gases – Emission, Measurement and Management

from fossil fuel varies with the fuel type and technology. Fossil fuel combustion and industrial processes are responsible for N2O emissions of around 0.7 Tg N per year (Denman, 2007). Other important sources include human sewage and the burning of

Across the agricultural landscape, several microbial processes can occur simultaneously to produce harmful N2O and non-greenhouse N2 in soils (pasture and arable) and sediments (drain, ditch, wetland and stream). These microbial processes are regulated by various soil, environmental and management factors, therefore making it difficult to control the rates of N2O and N2 production and their ratios (Paul & Beauchamp, 1989; Stevens et al., 1997; Zaman et al., 2007, 2008 b,c; Zaman & Nguyen, 2010). The aim of our review is to examine current information and understanding of the sources of N2O and N2 production and factors affecting N2O:N2 ratio in agricultural landscape, to enable management practices to be

Soil microbial processes, which account for major N2O production include; nitrification (Inubushi et al., 1996), denitrification (Tiedje, 1988; Firestone & Davidson, 1989; Smith, 1990;

al., 2001). These three microbial processes may occur in soils and sediments across the landscape depending on the physical (moisture contents or O2 level) and chemical

Autotrophic nitrification, a strictly aerobic process, is carried by chemolitho-autotrophic bacteria which use O2 as a terminal electron acceptor. In the first step, NH4+ is oxidized to NO2- by ammonia oxidizing species of the genus *Nitrosomonas,* while in the second step,

1981; Watson et al. 1981) as shown in Eq. 3.1.1 Other genera including *Nitrosococus, Nitrosospira* and subgenera *Nitrosobolus, and Nitrosovibrio* also have the ability to

(3.1.1)

In addition to NO2- production during the first stage of autotrophic nitrification, several intermediate and unstable compounds such as hydroxylamine (NH2OH) and nitroxyl (NOH) are also formed. Ammonia oxidizers consume relatively large amounts of molecular O2 during this first stage, causing anaerobic conditions in the microsites, which then lead to

Bock, 1998; Colliver & Stephenson, 2000) as shown in Eq. 3.1.2.


+ and NO3-), pH and C contents] in their micro-sites. Details of


+ (DNRA) (Silver et

devised that minimize N losses as N2O emission to the atmosphere.

Cavigelli & Robertson, 2001) and dissimilatory NO3- reduction to NH4

biomass and biofuels.

conditions [N form (i.e. NH4

**3.1 Nitrification** 

NO2- oxidation to NO3

a reduction of NO2

each of these processes are given below:

autotrophically oxidize NH3 to NO2-:

Denitrification is predominantly a microbial process by which NO3 - and NO2- are reduced to N2O and N2 in a respiratory metabolism. During respiratory denitrification, denitrifiers couple reduction of N-oxides to oxidation of organic C under anaerobic conditions and

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

slow-draining upland soils where anaerobic sites are prevalent.

N2O to N2.

**landscape** 

also been used to quantify N2O emissions.

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 11

transformation relative to denitrification under more reducing (O2 limited) conditions, since the microbes capable of DNRA are fermentative, and are able to grow in the absence of O2 contrary to predominantly aerobic denitrifiers. Silver et al. (2001) reported that in upland tropical forest soils, DNRA accounted for 75% of the turnover of the NO3- pool and N2O emission rates via DNRA, were 3 times greater than the combined N2O and N2 fluxes from nitrification and denitrification. Within the agricultural landscape, DNRA is likely to be an important N transformation process in wetland or stream sediments but may also occur in

(3.3.1)

As discussed above, while nitrification and DNRA produce only N2O, denitrification produces both N2O and N2. Stevens and Laughlin (1998) hypothesized that N2O produced by various processes might form one common pool before being reduced to N2 by nitrous oxide reductase. However, there is limited information available about the bulk reduction of

**4. Challenges in N2O measurements and estimates across the agricultural** 

Nitrous oxide emission and estimation across the different agricultural landscapes (arable, pasture, and wetland) is extremely variable (both spatially and temporally), thus posing the greatest challenge to researchers, modellers and policy makers to accurately predict N2O emissions. Among the different field and laboratory methods, the static chamber method has most widely been used to determine the rate of N2O emissions from soil because these chambers are easy to design, portable, compact, easy to install, and can be readily adapted to take gas measurements in the presence of animals and growing crops (Saggar et al., 2009). Readers are referred to Saggar et al (2009) for detailed information on the static chamber method. Other methods, including the sub-surface measurement of N2O emissions (Arah et al., 1991; Gut et al., 1998; Clark et al., 2001), the Push and Pull technique of Addy et al. (2002) modified by Zaman et al., 2008b to quantify N2O and N2 emissions from wetland soils and the estimation approach of the Intergovernmental Panel on Climate Change (IPCC) have

Few studies have carried out simultaneous measurements of N2 and N2O across the agricultural landscape. This is probably due to a lack of robust, easy and less expensive measurements and analytical methods. The most commonly used methods for measuring production of N2 and N2O in and their emission from the soils, include a technique based on the acetylene (C2H2) inhibition of N2O reduction (Tiedje et al., 1988) and methods using substrates enriched in 15N which allow subsequent 15N gas determination by isotope-ratio mass spectrometry (Mosier & Klemedtsson, 1994). These methods are not only expensive but far from perfect and have some serious biological implications. For example, C2H2 inhibition method needs paired soil samples (with or without C2H2), which is not only time

produce adenosine tri-phosphate (ATP) by phosphorylation (Firestone, 1982; Linn & Doran, 1984; Tiedje, 1988, Smith, 1990; Cavigelli & Robertson, 2001). Four different reductase enzymes are involved in a complete denitrification reaction. These enzymes are usually distributed in different microorganisms as shown in Eq. 3.2.1.

Denitrifiers are usually aerobic bacteria; however they prefer to use N-oxides at a low O2 level (Tiedje, 1988). Biological denitrification thus requires; NO3 - as a substrate (more than 2 mg NO3 --N per kg of soil) as an electron acceptor, absence of O2, which is related to a high soil moisture content >60% WFPS, available organic C as an electron donor, suitable soil pH, which generally ranges from 5 to 8 (optimum at 7) and a soil temperature range between 5 and 30 oC (optimum 25 oC) (Ryden & Lund, 1980; Ryden, 1983; Goodroad & Keeney, 1984; Scholefield et al., 1997; Barton et al., 1999; Swerts et al., 1997; Aulakh et al., 2001; Zaman et al., 2004; Zaman et al., 2007, 2008 b c, 2009; Zaman & Nguyen, 2010). However, the most critical factors are the NO3 concentrations, anaerobic conditions and the availability of soluble organic C (Zaman et al., 2007; 2008bc). Thus denitrification is expected to be an important N transformation process in areas where soils and sediments are subject to water logging (making them anaerobic), where they contain sufficient organic C and intercept inputs of NO3- or NO2- in groundwater or where there is excess nitrate after application of nitrate based fertilizers, or after nitrification (eg. 3.1.1). These areas include; urine patches in grazed pastures, where up to 1,000 kg N ha-1 can be found (Saggar et al., 2009; Zaman & Blennerhassett., 2010), riparian wetlands (Nguyen et al., 1999; Matheson et al., 2003), drains and ditches, and stream or river channels (Garcia-Ruiz et al., 1998; Bronson & Fillery, 1998; Mcmahon & Dennehy, 1999; Walker et al., 2002; Groffman et al., 2002; Zaman et al., 2008b&c, Zaman & Nguyen, 2010). However denitrification can also occur in less obviously waterlogged areas within the agricultural landscape due to the existence of anaerobic microsites such as in the center of soil aggregates (Parkin, 1987) or in areas of localized high O2 consumption (hot spots), which are created by decaying organic C (Burton et al., 1999; Godde & Conrad, 2000; Khalil et al., 2002; Mosier et al., 2002). Depending on soil physical and chemical conditions, other processes like chemo-denitrification can result in substantial production of N2O.

#### **3.3 Dissimilatory NO3 reduction to NH4 + (DNRA)**

DNRA is the 3rd biological process, which is known to produce considerable amounts of N2O as a byproduct under anaerobic conditions (Tiedje, 1988; Silver et al., 2001) as shown in Eq. 3.3.1

Conditions required for DNRA are similar to those required for denitrification and besides anaerobiosis include available NO3 - and organic C (Tiedje, 1988). DNRA has been found in anaerobic sludge and animal rumen, and also in lake littoral sediments, riparian wetland soil (Matheson et al., 2002) and tropical forest soils (Silver et al., 2001). Matheson et al., (2002) has also shown that DNRA is likely to be a more important process of NO3-

produce adenosine tri-phosphate (ATP) by phosphorylation (Firestone, 1982; Linn & Doran, 1984; Tiedje, 1988, Smith, 1990; Cavigelli & Robertson, 2001). Four different reductase enzymes are involved in a complete denitrification reaction. These enzymes are usually

(3.2.1)

Denitrifiers are usually aerobic bacteria; however they prefer to use N-oxides at a low O2 level (Tiedje, 1988). Biological denitrification thus requires; NO3- as a substrate (more than 2 mg NO3--N per kg of soil) as an electron acceptor, absence of O2, which is related to a high soil moisture content >60% WFPS, available organic C as an electron donor, suitable soil pH, which generally ranges from 5 to 8 (optimum at 7) and a soil temperature range between 5 and 30 oC (optimum 25 oC) (Ryden & Lund, 1980; Ryden, 1983; Goodroad & Keeney, 1984; Scholefield et al., 1997; Barton et al., 1999; Swerts et al., 1997; Aulakh et al., 2001; Zaman et al., 2004; Zaman et al., 2007, 2008 b c, 2009; Zaman & Nguyen, 2010). However, the most critical factors are the NO3- concentrations, anaerobic conditions and the availability of soluble organic C (Zaman et al., 2007; 2008bc). Thus denitrification is expected to be an important N transformation process in areas where soils and sediments are subject to water logging (making them anaerobic), where they contain sufficient organic C and intercept inputs of NO3- or NO2- in groundwater or where there is excess nitrate after application of nitrate based fertilizers, or after nitrification (eg. 3.1.1). These areas include; urine patches in grazed pastures, where up to 1,000 kg N ha-1 can be found (Saggar et al., 2009; Zaman & Blennerhassett., 2010), riparian wetlands (Nguyen et al., 1999; Matheson et al., 2003), drains and ditches, and stream or river channels (Garcia-Ruiz et al., 1998; Bronson & Fillery, 1998; Mcmahon & Dennehy, 1999; Walker et al., 2002; Groffman et al., 2002; Zaman et al., 2008b&c, Zaman & Nguyen, 2010). However denitrification can also occur in less obviously waterlogged areas within the agricultural landscape due to the existence of anaerobic microsites such as in the center of soil aggregates (Parkin, 1987) or in areas of localized high O2 consumption (hot spots), which are created by decaying organic C (Burton et al., 1999; Godde & Conrad, 2000; Khalil et al., 2002; Mosier et al., 2002). Depending on soil physical and chemical conditions, other processes like chemo-denitrification can result in substantial

nitric oxide reductase

ATP ATP ATP ATP

nitrous oxide reductase N2

distributed in different microorganisms as shown in Eq. 3.2.1.


NO3


production of N2O.

Eq. 3.3.1

**3.3 Dissimilatory NO3**

**-**

 **reduction to NH4**

**+ (DNRA)**  DNRA is the 3rd biological process, which is known to produce considerable amounts of N2O as a byproduct under anaerobic conditions (Tiedje, 1988; Silver et al., 2001) as shown in

Conditions required for DNRA are similar to those required for denitrification and besides anaerobiosis include available NO3- and organic C (Tiedje, 1988). DNRA has been found in anaerobic sludge and animal rumen, and also in lake littoral sediments, riparian wetland soil (Matheson et al., 2002) and tropical forest soils (Silver et al., 2001). Matheson et al., (2002) has also shown that DNRA is likely to be a more important process of NO3transformation relative to denitrification under more reducing (O2 limited) conditions, since the microbes capable of DNRA are fermentative, and are able to grow in the absence of O2 contrary to predominantly aerobic denitrifiers. Silver et al. (2001) reported that in upland tropical forest soils, DNRA accounted for 75% of the turnover of the NO3 - pool and N2O emission rates via DNRA, were 3 times greater than the combined N2O and N2 fluxes from nitrification and denitrification. Within the agricultural landscape, DNRA is likely to be an important N transformation process in wetland or stream sediments but may also occur in slow-draining upland soils where anaerobic sites are prevalent.

As discussed above, while nitrification and DNRA produce only N2O, denitrification produces both N2O and N2. Stevens and Laughlin (1998) hypothesized that N2O produced by various processes might form one common pool before being reduced to N2 by nitrous oxide reductase. However, there is limited information available about the bulk reduction of N2O to N2.

## **4. Challenges in N2O measurements and estimates across the agricultural landscape**

Nitrous oxide emission and estimation across the different agricultural landscapes (arable, pasture, and wetland) is extremely variable (both spatially and temporally), thus posing the greatest challenge to researchers, modellers and policy makers to accurately predict N2O emissions. Among the different field and laboratory methods, the static chamber method has most widely been used to determine the rate of N2O emissions from soil because these chambers are easy to design, portable, compact, easy to install, and can be readily adapted to take gas measurements in the presence of animals and growing crops (Saggar et al., 2009). Readers are referred to Saggar et al (2009) for detailed information on the static chamber method. Other methods, including the sub-surface measurement of N2O emissions (Arah et al., 1991; Gut et al., 1998; Clark et al., 2001), the Push and Pull technique of Addy et al. (2002) modified by Zaman et al., 2008b to quantify N2O and N2 emissions from wetland soils and the estimation approach of the Intergovernmental Panel on Climate Change (IPCC) have also been used to quantify N2O emissions.

Few studies have carried out simultaneous measurements of N2 and N2O across the agricultural landscape. This is probably due to a lack of robust, easy and less expensive measurements and analytical methods. The most commonly used methods for measuring production of N2 and N2O in and their emission from the soils, include a technique based on the acetylene (C2H2) inhibition of N2O reduction (Tiedje et al., 1988) and methods using substrates enriched in 15N which allow subsequent 15N gas determination by isotope-ratio mass spectrometry (Mosier & Klemedtsson, 1994). These methods are not only expensive but far from perfect and have some serious biological implications. For example, C2H2 inhibition method needs paired soil samples (with or without C2H2), which is not only time

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

N2O budget.

leachate and NH4

the higher C in riparian soils and enriched NO3

from streams, rivers and estuaries.

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 13

Lovell & Jarvis, 1996; Velthof et al., 1996; Jambert et al., 1997; Goossens, et al., 2001; De Klein et al., 2003**;** Zaman et al. 2007; 2008b, c; Saggar et al., 2007b; Zaman & Nguyen, 2010) which is greater than the 1% EF value of the IPCC. A comprehensive review collected by Saggar et al., (2009) indicated that N2O emissions from synthetic fertilizers range between 0.1 and about 2% of applied N. The large variations in the EF could be related to differences in soil types, time of fertilizer application, climatic conditions, weather patterns and form of synthetic fertilizers (ammonium and nitrate-based chemical fertilizers), animal urine and different protocols of N2O measurement such as static chambers, C2H2 inhibition, micrometeorological, and isotopic methods. Crutzen et al (2007) also reported that the IPCC methodology seriously underestimates N2O emissions from agriculture. Their estimates, using known global atmospheric removal rates and concentration growth of N2O, show an overall EF of 3–5%, whereas the EFs estimated for direct and indirect emissions using IPCC methodology cover only part of these emissions. Saggar et al (2009) further argued that the IPCC approach is limited by a number of uncertainties in emission factors, and in indirect emissions, limited data on the type and amount of N excreted by grazing animals, and in spatial and temporal variability of N2O emissions. Furthermore, the IPCC methodology does not allow for any mitigation options such as the use of urease or nitrification inhibitors and others. It is therefore critical to collect more data to validate the IPCC emission factor for N2O emission from agricultural soils, which may enable us to accurately predict the global

According to the IPCC, indirect N2O emission consists of 3 parts; N2O emissions associated with atmospheric N deposition [N2O (G)], human waste [N2O (S)], and with N lost via surface runoff and leaching [N2O (L)]. Indirect N2O emissions represent 1/3 of the total agricultural emissions, and the majority (75% of the total indirect emission) come from riparian zones (riparian wetlands, drainage ditch and stream sediments), where NO3- in

(Groffman, 2002; Zaman et al., 2007; Zaman et al., 2008b,c Zaman and Nguyen, 2010). N2O emission rates from riparian wetlands are generally higher than those of agricultural soils (Lowrance et al., 1984; Pinay et al., 1993; Zaman et al., 2008c) which could be attributed to

and groundwater flow to riparian zones. Given the potentially higher N2O emission rates from wet soils c.f. dry soils in agricultural landscapes, and the general lack of data from wet soils, there is a clear need for more data on N2O emission rates from riparian wetlands. Limited studies have been conducted to measure the rate of N2O emissions from streams and rivers. Garcia-Ruiz et al. (1999) found that N2 production ranged from 0.05-0.27 µmol N m-2 h-1 in the Swale-Ouse River system to 570 µmol N m-2 h-1 in the River Wiske. In the River Wiske, N2O production accounted for up to 80% of total N gas production. Using the current IPCC methodology, approximately 40% of indirect N2O emissions (emissions not accounted for from direct N sources such as fertilizers and applied animal urine) are derived

**5. Factors affecting N2O and N2 emissions and possible mitigating options**  As reviewed in Section 3 (Processes of N2O and N2 production across the agricultural landscape), autotrophic and heterotrophic nitrification and DNRA produce only N2O; while

+ in runoff from farmland are microbially converted to N2O and N2


consuming and expensive to analyze but a small amount of C2H2 (1%) can block nitrification and thus underestimates denitrification in NO3 - limited soils. Denitrifiers after repeated exposure to C2H2 adapt to C2H2 and use it as a source of C, which stimulates denitrification rates. Therefore both paired soil samples need to be discarded after 24 hrs of incubation to avoid this problem. In addition, acetone, which is added to C2H2 as stabilizer, also acts as a source of C for denitrifiers (Gross & Bremner, 1992) and needs to be removed before injecting C2H2 into soil cores or incubation jars. The most problematic step of this technique however, is to successfully achieve a uniform distribution of the desired concentration of C2H2 in microsites inhabited by denitrifiers if intact soils cores are used (Zaman & Nguyen, 2010). Similarly a lack of inhibitory effect of C2H2 on *Nitrosospira briensis*, one of the common ammonia-oxidizing bacteria in soils, observed by Wrage et al., (2004) also poses a challenge, especially in soil treated with ammonium-based fertilizer where N2O production via nitrifier-denitrification is likely to be overestimated. Thus although the technique of C2H2 inhibition has been widely used in laboratory conditions, when sieved soils or small monoliths were deployed, it has rarely been used in field conditions. To avoid the inhibitory effects of C2H2 on nitrification and denitrification, recently there has been an increasing interest in developing isotopic methods, which enable researchers to measure both N2O and N2 concurrently and identify the source of N2O production from various microbial processes including nitrification, denitrification and DNRA (Stevens et al., 1997; Matheson et al., 2003; Sutka et al., 2006). N2O production during nitrification and denitrification involves significant isotopic discrimination (ε = 3560‰ and 2833‰, respectively) (Robinson, 2001). Tilsner et al. (2003) reported that N2O emitted during denitrification under controlled laboratory conditions was highly depleted in 15N (40.8 ± 5.7‰). Similarly Stevens et al. (1997) differentially labelled the NH4+ and NO3 pools simultaneously with 15N, and periodically measured their individual 15N enrichments and N2O emission. A random distribution of 15N in N2O indicated a single source of origin whereas a non-random distribution indicated the two or more sources of N2O origin. Despite the fact that the isotopic method permits the fractional contribution of each pathway to N2O production and concurrent measurements of both N2O and N2, few researchers have used this method due to the high cost of 15N-substrates and 15N gases analyses, limited access to gas chromatograph with isotope-ratio mass spectrometers, and the difficulties associated with the uniform labeling of N pools in drier soils. Recently Mondini et al (2010) developed a robust automated dynamic closed chamber technique for concurrent measurement and analysis of N2O, CO2 and CH4 under laboratory conditions. In their system, a gas chromatograph is connected to a fully computerised sampling system composed of 16 sample jars and 2 multiposition valves. For further details on these various methods, the readers are referred to the above mentioned papers.

In the estimation approach, the IPCC divides N2O emissions from the agricultural landscape into direct and indirect emissions. Direct N2O emission refers to N2O derived from applied fertilizer and manure N, which is believed to increase with fertilizer use. Under the United Nations Framework Convention for Climate Change (UNFCCC), the majority of the countries use the IPCC default value of the 1% as emission factor (EF) (IPCC, 2006) from agricultural soils receiving synthetic fertilizers, farm dairy effluents (FDE), organic wastes and N fixed via biological fixation by leguminous crops (Bouwman et al., 2002; Stehfest & Bouwman 2006). However, a wide range of direct N2O emissions (i.e. 3 to 22% of applied N) across the agricultural landscape have been reported in the literature (Corre et al., 1996;

consuming and expensive to analyze but a small amount of C2H2 (1%) can block nitrification and thus underestimates denitrification in NO3- limited soils. Denitrifiers after repeated exposure to C2H2 adapt to C2H2 and use it as a source of C, which stimulates denitrification rates. Therefore both paired soil samples need to be discarded after 24 hrs of incubation to avoid this problem. In addition, acetone, which is added to C2H2 as stabilizer, also acts as a source of C for denitrifiers (Gross & Bremner, 1992) and needs to be removed before injecting C2H2 into soil cores or incubation jars. The most problematic step of this technique however, is to successfully achieve a uniform distribution of the desired concentration of C2H2 in microsites inhabited by denitrifiers if intact soils cores are used (Zaman & Nguyen, 2010). Similarly a lack of inhibitory effect of C2H2 on *Nitrosospira briensis*, one of the common ammonia-oxidizing bacteria in soils, observed by Wrage et al., (2004) also poses a challenge, especially in soil treated with ammonium-based fertilizer where N2O production via nitrifier-denitrification is likely to be overestimated. Thus although the technique of C2H2 inhibition has been widely used in laboratory conditions, when sieved soils or small monoliths were deployed, it has rarely been used in field conditions. To avoid the inhibitory effects of C2H2 on nitrification and denitrification, recently there has been an increasing interest in developing isotopic methods, which enable researchers to measure both N2O and N2 concurrently and identify the source of N2O production from various microbial processes including nitrification, denitrification and DNRA (Stevens et al., 1997; Matheson et al., 2003; Sutka et al., 2006). N2O production during nitrification and denitrification involves significant isotopic discrimination (ε = 3560‰ and 2833‰, respectively) (Robinson, 2001). Tilsner et al. (2003) reported that N2O emitted during denitrification under controlled laboratory conditions was highly depleted in 15N (40.8 ± 5.7‰). Similarly Stevens et al.

+ and NO3

periodically measured their individual 15N enrichments and N2O emission. A random distribution of 15N in N2O indicated a single source of origin whereas a non-random distribution indicated the two or more sources of N2O origin. Despite the fact that the isotopic method permits the fractional contribution of each pathway to N2O production and concurrent measurements of both N2O and N2, few researchers have used this method due to the high cost of 15N-substrates and 15N gases analyses, limited access to gas chromatograph with isotope-ratio mass spectrometers, and the difficulties associated with the uniform labeling of N pools in drier soils. Recently Mondini et al (2010) developed a robust automated dynamic closed chamber technique for concurrent measurement and analysis of N2O, CO2 and CH4 under laboratory conditions. In their system, a gas chromatograph is connected to a fully computerised sampling system composed of 16 sample jars and 2 multiposition valves. For further details on these various methods, the

In the estimation approach, the IPCC divides N2O emissions from the agricultural landscape into direct and indirect emissions. Direct N2O emission refers to N2O derived from applied fertilizer and manure N, which is believed to increase with fertilizer use. Under the United Nations Framework Convention for Climate Change (UNFCCC), the majority of the countries use the IPCC default value of the 1% as emission factor (EF) (IPCC, 2006) from agricultural soils receiving synthetic fertilizers, farm dairy effluents (FDE), organic wastes and N fixed via biological fixation by leguminous crops (Bouwman et al., 2002; Stehfest & Bouwman 2006). However, a wide range of direct N2O emissions (i.e. 3 to 22% of applied N) across the agricultural landscape have been reported in the literature (Corre et al., 1996;

pools simultaneously with 15N, and

(1997) differentially labelled the NH4

readers are referred to the above mentioned papers.

Lovell & Jarvis, 1996; Velthof et al., 1996; Jambert et al., 1997; Goossens, et al., 2001; De Klein et al., 2003**;** Zaman et al. 2007; 2008b, c; Saggar et al., 2007b; Zaman & Nguyen, 2010) which is greater than the 1% EF value of the IPCC. A comprehensive review collected by Saggar et al., (2009) indicated that N2O emissions from synthetic fertilizers range between 0.1 and about 2% of applied N. The large variations in the EF could be related to differences in soil types, time of fertilizer application, climatic conditions, weather patterns and form of synthetic fertilizers (ammonium and nitrate-based chemical fertilizers), animal urine and different protocols of N2O measurement such as static chambers, C2H2 inhibition, micrometeorological, and isotopic methods. Crutzen et al (2007) also reported that the IPCC methodology seriously underestimates N2O emissions from agriculture. Their estimates, using known global atmospheric removal rates and concentration growth of N2O, show an overall EF of 3–5%, whereas the EFs estimated for direct and indirect emissions using IPCC methodology cover only part of these emissions. Saggar et al (2009) further argued that the IPCC approach is limited by a number of uncertainties in emission factors, and in indirect emissions, limited data on the type and amount of N excreted by grazing animals, and in spatial and temporal variability of N2O emissions. Furthermore, the IPCC methodology does not allow for any mitigation options such as the use of urease or nitrification inhibitors and others. It is therefore critical to collect more data to validate the IPCC emission factor for N2O emission from agricultural soils, which may enable us to accurately predict the global N2O budget.

According to the IPCC, indirect N2O emission consists of 3 parts; N2O emissions associated with atmospheric N deposition [N2O (G)], human waste [N2O (S)], and with N lost via surface runoff and leaching [N2O (L)]. Indirect N2O emissions represent 1/3 of the total agricultural emissions, and the majority (75% of the total indirect emission) come from riparian zones (riparian wetlands, drainage ditch and stream sediments), where NO3- in leachate and NH4+ in runoff from farmland are microbially converted to N2O and N2 (Groffman, 2002; Zaman et al., 2007; Zaman et al., 2008b,c Zaman and Nguyen, 2010). N2O emission rates from riparian wetlands are generally higher than those of agricultural soils (Lowrance et al., 1984; Pinay et al., 1993; Zaman et al., 2008c) which could be attributed to the higher C in riparian soils and enriched NO3- inputs from surrounding areas via seepage and groundwater flow to riparian zones. Given the potentially higher N2O emission rates from wet soils c.f. dry soils in agricultural landscapes, and the general lack of data from wet soils, there is a clear need for more data on N2O emission rates from riparian wetlands. Limited studies have been conducted to measure the rate of N2O emissions from streams and rivers. Garcia-Ruiz et al. (1999) found that N2 production ranged from 0.05-0.27 µmol N m-2 h-1 in the Swale-Ouse River system to 570 µmol N m-2 h-1 in the River Wiske. In the River Wiske, N2O production accounted for up to 80% of total N gas production. Using the current IPCC methodology, approximately 40% of indirect N2O emissions (emissions not accounted for from direct N sources such as fertilizers and applied animal urine) are derived from streams, rivers and estuaries.

## **5. Factors affecting N2O and N2 emissions and possible mitigating options**

As reviewed in Section 3 (Processes of N2O and N2 production across the agricultural landscape), autotrophic and heterotrophic nitrification and DNRA produce only N2O; while

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

 **concentrations** 

**-**

The amount of mineral N, both NH4+ and NO3-

**5.1 Soil NH4**

**+**

entry of urea-N into the NH4

(Eq. 3.1.2).

 **and NO3**

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 15

and their ratio. The amount of NH4+ in soil under aerobic soil conditions, and hence its availability for nitrification, can directly regulate N2O emission via nitrifier-denitrification (Webster & Hopkins, 1996; Wrage et al., 2001; Dalal et al., 2003; Ma et al., 2007), while NO3 is used as a substrate by both denitrification and DNRA and thus affects N2O production **(**Webster and Hopkins, 1996; Zaman et al., 2008c). A higher level of NO3- in soil is also known to result in incomplete denitrification and thus higher N2O:N2 due to suppression of nitrous oxide reductase activity, the enzyme responsible for microbial conversion of N2O to N2 (Eq. 3.2.1). To mitigate N2O emissions, researchers during the past two decades focused mainly on reducing the rate of nitrification while little work has been done to exert control on the denitrification level. For example, to reduce N2O emissions from applied urea, ammonium based fertilizers or urine N, researchers have developed different mitigation technologies including the use of N inhibitors to reduce the entry of mineral N from applied fertilizer/urine into the available N pool, application of soil amendments like zeolites to capture soil NH4+ and controlled release and split applications of N fertilizers to match crop N demand. Among these options, coating chemical fertilizers with N inhibitors or applying N inhibitors on their own to treat urine patches in grazed pastures have received the most attention. The two major classes of N inhibitors are urease inhibitors (UIs) and nitrification inhibitors (NIs). Urease inhibitors retard the hydrolysis of soil-applied urea and delay the

to the limited availability of NH4+ (Watson, 2000; Xu et al., 2000; Zaman et al., 2008a; Zaman et al., 2009) as shown earlier in Eq. 2.1.4. Such a reduction in urea hydrolysis also limits the opportunity for nitrite (NO2-) accumulation in the soil, which is known to produce N2O

Decisions about N fertilizer application are usually dependent on the availability of water, and the N application rate is determined by crop growth stage and the productivity goals. Fast urea hydrolysis starts within hours of urea fertilizer application or after deposition of urine from grazing animals and is completed within 1 to 3 days, during which time a significant amount of NH3 (up to 30% of the applied N) is lost. UI like N-(n-butyl) thiophosphoric triamide (nBTPT) or Agrotain® applied at a very low concentration (0.01%) with urea fertilizer is reported to delay such fast urea hydrolysis by 7 to 9 days (Watson et al., 2008), which has implications for worldwide urea use in pasture and cropping systems where there is a high risk of NH3 loss due to low moisture and high temperature, especially during summer or early autumn. Such a delay in urea hydrolysis allows more time for rainfall or irrigation water to wash the applied urea from surface soil and thus minimizes the risk of NH3 emissions. After application, nBTPT is quickly converted to its oxygen analog N-(n-butyl) phosphoric triamide) (NBPT) (Eq. 5.1.1), which is the actual UI (McCarty et al., 1989; Christianson et al., 1990; Creason et al., 1990), and it is bound and moves along with urea molecule in the soil (Christianson & Howard, 1994). The conversion of nBTPT to NBPT is rapid, occurring within minutes/hours in aerobic soils (Byrnes & Freney, 1995), but can take several days in the floodwater of tropical soils. NBPT forms a tridentate ligand with

In addition to reduced nitrification, Agrotain is also known to reduce N2O emission indirectly through reduced NH3 volatilization (Watson et al., 1990; 1994 a & b, 1998, 2008;

the urease enzyme, blocking the active site (Manunza et al., 1999).

, are critical for the production of N2O, N2

+ pool, which is likely to produce less N2O via nitrification due

denitrification produces both N2O and N2. The emissions of N2O and N2 and their ratios are affected by various soil and management factors, including mineral N concentration, available C, soil pH, soil aeration status, soil temperature and their interactions as shown in Table 5.1.


Table 5.1. Factors affecting N2O and N2 emissions and their ratios.

In the section below, an attempt has been made to discuss these soil and management factors. Understanding these factors may help us to design mitigating tools to reduce the rate of N2O production and to lower N2O:N2 ratios.

#### **5.1 Soil NH4 + and NO3 concentrations**

14 Greenhouse Gases – Emission, Measurement and Management

denitrification produces both N2O and N2. The emissions of N2O and N2 and their ratios are affected by various soil and management factors, including mineral N concentration, available C, soil pH, soil aeration status, soil temperature and their interactions as shown in

Factors Management Impact Management practices

denitrification and lower N2O:N2

Use urease and

nitrification inhibitors to enhance fertilizer use efficiency; split N fertilizer application to synchronize plant N demand and to minimize N losses; avoid over grazing; manipulation of animal diet; use of constructed or natural riparian wetland, improving water use efficiency to avoid anaerobicity

Sequester more C by adopting management practices including zero or minimum tillage, retention of crop residues, mulching, application of organic and farm wastes, biochar, and applying chemical fertilizers with organic

amendments

use efficiency

Regular liming each year or if possible with every N fertilizer application

Improving soil structure via C sequestration, avoiding soil compaction; improving soil drainage condition and also water

Decrease Reduce nitrification &

facilitate

denitrification and thus lower N2O:N2

and denitrification and thus lower

and denitrification and thus lower

In the section below, an attempt has been made to discuss these soil and management factors. Understanding these factors may help us to design mitigating tools to reduce the

N2O:N2

N2O:N2

Improve Facilitate nitrification

Soil organic C Increase Improve soil health,

Soil pH Increase Facilitate nitrification

Table 5.1. Factors affecting N2O and N2 emissions and their ratios.

rate of N2O production and to lower N2O:N2 ratios.

Table 5.1.

Soil NH4

concentration

Soil aeration and water status

+ & NO3-

The amount of mineral N, both NH4+ and NO3- , are critical for the production of N2O, N2 and their ratio. The amount of NH4 + in soil under aerobic soil conditions, and hence its availability for nitrification, can directly regulate N2O emission via nitrifier-denitrification (Webster & Hopkins, 1996; Wrage et al., 2001; Dalal et al., 2003; Ma et al., 2007), while NO3 is used as a substrate by both denitrification and DNRA and thus affects N2O production **(**Webster and Hopkins, 1996; Zaman et al., 2008c). A higher level of NO3 - in soil is also known to result in incomplete denitrification and thus higher N2O:N2 due to suppression of nitrous oxide reductase activity, the enzyme responsible for microbial conversion of N2O to N2 (Eq. 3.2.1). To mitigate N2O emissions, researchers during the past two decades focused mainly on reducing the rate of nitrification while little work has been done to exert control on the denitrification level. For example, to reduce N2O emissions from applied urea, ammonium based fertilizers or urine N, researchers have developed different mitigation technologies including the use of N inhibitors to reduce the entry of mineral N from applied fertilizer/urine into the available N pool, application of soil amendments like zeolites to capture soil NH4 + and controlled release and split applications of N fertilizers to match crop N demand. Among these options, coating chemical fertilizers with N inhibitors or applying N inhibitors on their own to treat urine patches in grazed pastures have received the most attention. The two major classes of N inhibitors are urease inhibitors (UIs) and nitrification inhibitors (NIs). Urease inhibitors retard the hydrolysis of soil-applied urea and delay the entry of urea-N into the NH4 + pool, which is likely to produce less N2O via nitrification due to the limited availability of NH4+ (Watson, 2000; Xu et al., 2000; Zaman et al., 2008a; Zaman et al., 2009) as shown earlier in Eq. 2.1.4. Such a reduction in urea hydrolysis also limits the opportunity for nitrite (NO2-) accumulation in the soil, which is known to produce N2O (Eq. 3.1.2).

Decisions about N fertilizer application are usually dependent on the availability of water, and the N application rate is determined by crop growth stage and the productivity goals. Fast urea hydrolysis starts within hours of urea fertilizer application or after deposition of urine from grazing animals and is completed within 1 to 3 days, during which time a significant amount of NH3 (up to 30% of the applied N) is lost. UI like N-(n-butyl) thiophosphoric triamide (nBTPT) or Agrotain® applied at a very low concentration (0.01%) with urea fertilizer is reported to delay such fast urea hydrolysis by 7 to 9 days (Watson et al., 2008), which has implications for worldwide urea use in pasture and cropping systems where there is a high risk of NH3 loss due to low moisture and high temperature, especially during summer or early autumn. Such a delay in urea hydrolysis allows more time for rainfall or irrigation water to wash the applied urea from surface soil and thus minimizes the risk of NH3 emissions. After application, nBTPT is quickly converted to its oxygen analog N-(n-butyl) phosphoric triamide) (NBPT) (Eq. 5.1.1), which is the actual UI (McCarty et al., 1989; Christianson et al., 1990; Creason et al., 1990), and it is bound and moves along with urea molecule in the soil (Christianson & Howard, 1994). The conversion of nBTPT to NBPT is rapid, occurring within minutes/hours in aerobic soils (Byrnes & Freney, 1995), but can take several days in the floodwater of tropical soils. NBPT forms a tridentate ligand with the urease enzyme, blocking the active site (Manunza et al., 1999).

In addition to reduced nitrification, Agrotain is also known to reduce N2O emission indirectly through reduced NH3 volatilization (Watson et al., 1990; 1994 a & b, 1998, 2008;

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

DCD (Weiske et al., 2001; Zerulla et al., 2001).

**5.2 Soil available organic C concentration** 

fertilized soils, NO3-

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 17

water, non-volatile, cheap and can be easily treated/coated onto solid ammonium based N fertilizers such as urea, diammonium phosphate (DAP); ammonium nitrate (NH4NO3) and ammonium sulfate (NH4)2SO4 or directly added into FDE to improve their N use efficiency and minimize N losses. However after application, separation of DCD from applied NH4+, DCD leaching, and its rapid decomposition with increasing soil temperature are reported to lower its efficacy. Contrary to this, DMPP has several advantages over DCD and nitrapyrin. Lower application rates (0.5 to 1.0 kg of the active compound ha<sup>−</sup>1) are needed to achieve

application, DMPP is less prone to leaching and remains effective much longer than that of

In intensive agricultural system like grazed pastures, other mitigation options including feeding dairy cows with low N feed such as palm kernel and maize silage instead of high N pastures to reduce the amount of N in animal excreta, using winter feed pads and restricted grazing to avoid soil compaction and to minimize urine depositions during critical times (winter) (de Klein et al., 2006), natural and constructed riparian wetlands to intercept N entering from adjacent pasture soils and to process it before entering water bodies (Zaman et al., 2008b), applying lime or zeolite as soil amendments to reduce N2O emissions and shift the balance between harmful N2O and non-greenhouse N2 (Zaman et al., 2007, 2008c; Zaman & Nguyen, 2010), adding salts to animal feed to increase urine volume and spread (Ledgard et al., 2007), increasing the hippuric acid concentration in urine by manipulating animal feed

Soil organic C is another important controller of N2O and N2 production in soils and sediments as denitrifiers are strictly heterotrophs and use available organic C as electron donor and indirectly affects O2 concentrations of aerobic soils (Groffman et al., 1987). However the effect of available C on the amounts of N2O and N2 produced in and emitted from the soils, as well as on the ratio between the two gases, is reported to vary with soil NO3- concentration and WFPS (Zaman et al., 2007, 2008b,c). In anaerobic zones of non-

section 5.1, while in soils with high NO3- inputs (i.e. after application of chemical fertilizers and FDE or urine patches after grazing), available soil organic C would be the main driver of N2O and N2 production via denitrification (Tiedje, 1988). Applying urea fertilizer with C source (wheat straw and green manure) was reported to substantially reduce N2O emission compared to urea fertilizer alone (Aulakh et al., 2001) possibly due to the microbial immobilization (Tiedje, 1988) or DNRA (Matheson et al., 2002). Zaman et al., (2008b) during an incubation study observed that wetland soils treated with KNO3 emitted more N2 emissions than those of the pasture soils which they attributed to the availability of highly enriched organic C and high WFPS in wetland soils. Weier et al. (1993) also measured N2O and N2 emissions from 4 soils treated with a range of available C (0, 180 and 360 kg ha-1), NO3--N (0, 50 and 100 kg ha-1) and WFPS (60, 75 and 90%). They reported that N2 emission was favored at the highest available C rate of 360 kg C ha-1 and 90%WFPS, while the higher NO3- concentration inhibited the conversion of N2O to N2, resulting in higher N2O:N2 ratios. Similarly Yao et al. (2002) observed a negative correlation between N2O emission and soil organic C from N fertilized wheat crop. The N2O:N2 ratio could be explained by an

availability may control the denitrification rates as discussed above in

leaching. After

optimal nitrification inhibition to reduce N2O emissions and NO3-

(Bertram et al., 2009) have been suggested as additional mitigation tools.

(5.1.1)

Chadwick et al., 2005; Meneer et al., 2008; Sanz-Cobena et al., 2008; Singh et al., 2008; Zaman et al., 2008a & 2009; Zaman & Blennerhassett., 2010). Ammonia itself is not a greenhouse gas, but it can act as a secondary source of N2O production after its deposition on land (Martikainen, 1985) and thus contributes indirectly to climate change. To our knowledge, New Zealand is the only country that has included NH3 reduction from Agrotain treated urea in its national inventory on N2O. Sherlock et al. (2008) after a literature search on NH3 emission, recommended to the New Zealand Ministry of Agriculture and Forestry (MAF) that a specific value of 0.1 for FracGASM and FracGASF be considered for adoption. In New Zealand, studies where nBTPT (0.025% w/w) was applied reduced NH3 emissions by 43% from urea and by 48% from urine (Singh et al., 2008; Meneer et al., 2008; Zaman et al., 2008a & 2009; Zaman & Blennerhassett 2010). Based on these estimates of reductions in NH3 emission from nBTPT treated urea applications, a New Zealand specific value of 0.06 for FracGASF is recommended for adoption where fertilizers containing the urease inhibitor, nBTPT are applied. Saggar et al. (2010) recommended to MAF that where NBTPT is applied with urea, FracGASF should be calculated as follows,

$$\text{FracGASE} = \left[ \left( \text{FN}\_{\text{UI}} \right) \times 0.06 \right] + \left[ \left( \text{FN} - \text{FN}\_{\text{UI}} \right) \times 0.10 \right] \text{ }$$

Where **FracGASF** is the fraction of total fertilizer N emitted as NH3, **FN** is the total amount of applied fertilizer N, **FNUI** is the amount of applied fertilizer N treated with the urease inhibitor, nBTPT. Changing the FracGASF from 0.1 to 0.06 for the current use of 18.4 Gg N of Agrotain treated urea reduces indirect N2O emissions by 0.012 Gg, which equates to 3.6 Gg CO2-equiv. However, assuming all urea is applied with nBTPT in New Zealand, changing the FracGASF from 0.1 to 0.06 will reduce the indirect N2O emissions by 0.14 Gg, which equates to 43.4 Gg CO2-equiv.

Nitrification inhibitors are compounds (both natural and synthetic) that delay bacterial oxidation of NH4+ either by temporarily suppressing the activities of nitrifiers or killing them in the soil, thus maintaining the applied N in more stable form (i.e. NH4+-N). Slowing down nitrification in soils lowers N2O production associated with nitrifier-denitrification (Webster & Hopkins, 1996; Wrage et al., 2001; Ma et al., 2007), or indirectly by reducing the amount of NO3- substrate available for denitrification. Among the many synthetic NIs, only nitrapyrin or N-Serve (NP) (2-chloro-6-(tri-chloromethyl) pyridine), dicyandiamide (DCD) and 3,4-Dimethylpyrazol-phosphate (DMPP) have gained substantial practical and commercial importance in agricultural and horticultural crop production.

Nitrapyrin because of its high volatility needs to be injected into the soil. Therefore nitrapyrin may be a preferred NI where injecting chemical fertilizers or farm dairy effluent (FDE) into the soil is a common practice. Unlike nitrapyrin, DCD is relatively soluble in water, non-volatile, cheap and can be easily treated/coated onto solid ammonium based N fertilizers such as urea, diammonium phosphate (DAP); ammonium nitrate (NH4NO3) and ammonium sulfate (NH4)2SO4 or directly added into FDE to improve their N use efficiency and minimize N losses. However after application, separation of DCD from applied NH4+, DCD leaching, and its rapid decomposition with increasing soil temperature are reported to lower its efficacy. Contrary to this, DMPP has several advantages over DCD and nitrapyrin.

Lower application rates (0.5 to 1.0 kg of the active compound ha<sup>−</sup>1) are needed to achieve optimal nitrification inhibition to reduce N2O emissions and NO3 leaching. After application, DMPP is less prone to leaching and remains effective much longer than that of DCD (Weiske et al., 2001; Zerulla et al., 2001).

In intensive agricultural system like grazed pastures, other mitigation options including feeding dairy cows with low N feed such as palm kernel and maize silage instead of high N pastures to reduce the amount of N in animal excreta, using winter feed pads and restricted grazing to avoid soil compaction and to minimize urine depositions during critical times (winter) (de Klein et al., 2006), natural and constructed riparian wetlands to intercept N entering from adjacent pasture soils and to process it before entering water bodies (Zaman et al., 2008b), applying lime or zeolite as soil amendments to reduce N2O emissions and shift the balance between harmful N2O and non-greenhouse N2 (Zaman et al., 2007, 2008c; Zaman & Nguyen, 2010), adding salts to animal feed to increase urine volume and spread (Ledgard et al., 2007), increasing the hippuric acid concentration in urine by manipulating animal feed (Bertram et al., 2009) have been suggested as additional mitigation tools.

## **5.2 Soil available organic C concentration**

16 Greenhouse Gases – Emission, Measurement and Management

 (5.1.1)

Chadwick et al., 2005; Meneer et al., 2008; Sanz-Cobena et al., 2008; Singh et al., 2008; Zaman et al., 2008a & 2009; Zaman & Blennerhassett., 2010). Ammonia itself is not a greenhouse gas, but it can act as a secondary source of N2O production after its deposition on land (Martikainen, 1985) and thus contributes indirectly to climate change. To our knowledge, New Zealand is the only country that has included NH3 reduction from Agrotain treated urea in its national inventory on N2O. Sherlock et al. (2008) after a literature search on NH3 emission, recommended to the New Zealand Ministry of Agriculture and Forestry (MAF) that a specific value of 0.1 for FracGASM and FracGASF be considered for adoption. In New Zealand, studies where nBTPT (0.025% w/w) was applied reduced NH3 emissions by 43% from urea and by 48% from urine (Singh et al., 2008; Meneer et al., 2008; Zaman et al., 2008a & 2009; Zaman & Blennerhassett 2010). Based on these estimates of reductions in NH3 emission from nBTPT treated urea applications, a New Zealand specific value of 0.06 for FracGASF is recommended for adoption where fertilizers containing the urease inhibitor, nBTPT are applied. Saggar et al. (2010) recommended to MAF that where NBTPT is applied

FracGASF = [(FNUI) × 0.06] + [(FN– FNUI) × 0.10] Where **FracGASF** is the fraction of total fertilizer N emitted as NH3, **FN** is the total amount of applied fertilizer N, **FNUI** is the amount of applied fertilizer N treated with the urease inhibitor, nBTPT. Changing the FracGASF from 0.1 to 0.06 for the current use of 18.4 Gg N of Agrotain treated urea reduces indirect N2O emissions by 0.012 Gg, which equates to 3.6 Gg CO2-equiv. However, assuming all urea is applied with nBTPT in New Zealand, changing the FracGASF from 0.1 to 0.06 will reduce the indirect N2O emissions by 0.14 Gg, which equates to

Nitrification inhibitors are compounds (both natural and synthetic) that delay bacterial oxidation of NH4+ either by temporarily suppressing the activities of nitrifiers or killing them in the soil, thus maintaining the applied N in more stable form (i.e. NH4+-N). Slowing down nitrification in soils lowers N2O production associated with nitrifier-denitrification (Webster & Hopkins, 1996; Wrage et al., 2001; Ma et al., 2007), or indirectly by reducing the amount of NO3- substrate available for denitrification. Among the many synthetic NIs, only nitrapyrin or N-Serve (NP) (2-chloro-6-(tri-chloromethyl) pyridine), dicyandiamide (DCD) and 3,4-Dimethylpyrazol-phosphate (DMPP) have gained substantial practical and

Nitrapyrin because of its high volatility needs to be injected into the soil. Therefore nitrapyrin may be a preferred NI where injecting chemical fertilizers or farm dairy effluent (FDE) into the soil is a common practice. Unlike nitrapyrin, DCD is relatively soluble in

commercial importance in agricultural and horticultural crop production.

with urea, FracGASF should be calculated as follows,

43.4 Gg CO2-equiv.

Soil organic C is another important controller of N2O and N2 production in soils and sediments as denitrifiers are strictly heterotrophs and use available organic C as electron donor and indirectly affects O2 concentrations of aerobic soils (Groffman et al., 1987). However the effect of available C on the amounts of N2O and N2 produced in and emitted from the soils, as well as on the ratio between the two gases, is reported to vary with soil NO3- concentration and WFPS (Zaman et al., 2007, 2008b,c). In anaerobic zones of nonfertilized soils, NO3- availability may control the denitrification rates as discussed above in section 5.1, while in soils with high NO3- inputs (i.e. after application of chemical fertilizers and FDE or urine patches after grazing), available soil organic C would be the main driver of N2O and N2 production via denitrification (Tiedje, 1988). Applying urea fertilizer with C source (wheat straw and green manure) was reported to substantially reduce N2O emission compared to urea fertilizer alone (Aulakh et al., 2001) possibly due to the microbial immobilization (Tiedje, 1988) or DNRA (Matheson et al., 2002). Zaman et al., (2008b) during an incubation study observed that wetland soils treated with KNO3 emitted more N2 emissions than those of the pasture soils which they attributed to the availability of highly enriched organic C and high WFPS in wetland soils. Weier et al. (1993) also measured N2O and N2 emissions from 4 soils treated with a range of available C (0, 180 and 360 kg ha-1), NO3--N (0, 50 and 100 kg ha-1) and WFPS (60, 75 and 90%). They reported that N2 emission was favored at the highest available C rate of 360 kg C ha-1 and 90%WFPS, while the higher NO3- concentration inhibited the conversion of N2O to N2, resulting in higher N2O:N2 ratios. Similarly Yao et al. (2002) observed a negative correlation between N2O emission and soil organic C from N fertilized wheat crop. The N2O:N2 ratio could be explained by an

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

al., 2007, 2008b, Zaman & Nguyen, 2010).

**5.4 Soil aeration and water status** 

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 19

are reported to lower soil pH after their application (Thornton et al., 1996; Mulvaney et al., 1997; Nobre, 2001; Cai et al., 2002). For example, Mulvaney et al. (1997) have reported that ammonium-based fertilizers with soil acidifying effects produce a higher N2O:N2 ratio compared to alkaline forming fertilizers (anhydrous ammonia, urea or di-ammonium phosphate). Most researchers attribute high N2O and low N2 emissions in acidic conditions to the suppression of nitrous oxide-reductase at low soil pH (inhibition at soil pH 4.5) (Kostina et al., 1996; Daum & Schenk 1998; Flessa et al., 1998; Stevens and Laughlin, 1998; Zaman et al., 2007). It is also likely that all denitrifying enzymes are susceptible at low soil pH and produce N2O from intermediate products (Nagele & Conrad, 1990). However, the extensive review conducted by Šimek and Cooper (2002) reported that the lower rates of N2 and high N2O:N2 ratio at low soil pH could be due to lower amounts of soil organic C and mineral N available to the denitrifying population under acid conditions rather than a direct effect of low pH on denitrification enzymes. Regardless of the biochemical reasons for changes in soil pH on N2 emission, raising soil pH through application of soil amendment like lime appears to offer a mechanism for mitigation of N2O (Šimek et al., 2002; Zaman et

Soil aeration, namely O2 concentration and gas exchange between soil and atmosphere, affects all microbial N transformation processes including nitrification, denitrification, and DNRA. Control of the denitrification enzymes, especially nitrous oxide reductase, represents the key mitigation option for the rate of N2O production and can therefore shift the balance between harmful N2O and non-greenhouse N2 production across the agricultural landscape (Smith & Tiedje, 1979; Mosier et al., 1986; Robertson & Tiedje, 1987; Henrich & Haselwandter, 1997; Bollmann & Conrad, 1998; Mosier et al., 2002).). In soil, O2 concentration changes with soil moisture content and organic matter decomposition by soil microorganisms. After rainfall or applying irrigation water, soils become temporarily anaerobic; the extent and duration of anaerobiosis differs with soil types (drainage class). Fine-textured soils with a higher clay content are reported to remain anaerobic for a longer period of time at low WFPS than coarse-textured soils because of the greater number of micro pores in the former (Barton et al., 1999). Therefore fine-textured soils with poorly drained conditions are likely to emit more N2O for a longer period than those of coarsetextured soils with well-drained conditions (Groffman & Tiedje, 1989; Aulakh et al., 1991; Clayton et al., 1997; Dobbie & Smith, 2001; Saggar et al., 2004a). At higher O2, partial pressure (>0.5 vol. %), nitrification is expected to proceed, provided there is sufficient water for optimum activity of nitrifiers (Linn & Doran, 1984; Bollmann & Conrad, 1998); if the soil WFPS increases (and pO2 decreases), the rate of N2O production and the proportion of N2O to NO3- produced also increases (Smith et al., 2003). Under such specific conditions at WFPS>60%, nitrification is considered to be the predominant source of N2O as opposed to denitrification or DNRA (Inubushi et al., 1996). Although DNRA is understood to be an anaerobic process, information about the critical levels of WFPS or O2 for DNRA is lacking in the literature. Denitrification becomes a major source of N2O and N2 production at lower O2 partial pressure (<0.5 vol. %) and higher WFPS (>60%) (Davidson, 1993; Scholefield et al., 1997; Bronson & Fillery, 1998; Khalil et al., 2002). In such scenarios, more aerobic soils are likely to produce mainly N2O because denitrification reductases (Eq. 3.2.1) especially nitrous oxide reductase is reported to be sensitive to soil O2 level, while anaerobic soils and

interaction of C availability, NO3 concentration and enzyme status (Swerts et al., 1996). There are reports that the N2O:N2 ratios are lower in the rhizosphere, which provides more available organic C in the form of root exudates and root debris, and is characterized by low partial pressure of O2 (due to O2 consumption by plant roots) (Casella et al., 1984).

Depending on the management practices, agricultural soils can act as a source or sink for atmospheric CO2. Improved land management practices in croplands and grasslands can store up to 1 Gt C in the soil on an annual basis (IPCC, 2000). It is therefore possible to store between 100 to 1000 kg SOC ha-1 year-1 depending on the climate, soil and vegetation types, and site-specific soil management practices. Improved land management practices like zero or minimum tillage, retention of crop residues via crop rotation and mulching, application of FDE, organic residues and manure, following crop rotation especially with N fixing crops and avoiding burning crop residues after harvest may offer potential mitigation tool to sequester more C in the soils to offset the increase in atmospheric CO2 as well as to improve soil fertility, soil structure, aggregate stability, pore size geometry and distribution, water and nutrients holding capacity and soil quality (increased microbial and enzymatic activities). Such improvement in soil physical and chemical fertility and health will minimize conducive conditions like anaerobicity and soil compaction which stimulate denitrification. Increased soil C may also help to shift the balance between harmful N2O and non-greenhouse gas N2 during denitrification as the activity of nitrous oxide reductase enzymes is stimulated by available soil C.

#### **5.3 Soil pH**

Soil pH is among the key regulators of the microbiological processes that affect N2O and N2 production and their ratios. Nitrification activity is generally higher with higher soil pH (> 6) (Bremner & Blackmer, 1981; Bramley & White, 1989). The critical soil pH threshold for nitrification is 5; however, nitrification can occur even at a soil pH of 4.5 due to acid-adapted nitrifier strains (Bouwman, 1990). Denitrification has been reported to occur over a wide range of soil pH values (5 to 8) (Weier & Gilliam, 1986; Ramos, 1996; Flessa et al, 1998); however, laboratory experiments with artificially adjusted soil pH suggest, that under optimized conditions (very low pO2, NO3 and glucose amendments), denitrification can proceed even at pHs below 4 or above 10 (Šimek & Hopkins, 1999, Šimek et al., 2002). Numerous laboratory and field studies have shown that soil pH affects N2O and N2 and the ratio of these gases (e.g. Weier & Gilliam, 1986; Stevens & Laughlin, 1998). Under controlled environment experiments, we found that raising soil pH to 7 through lime application significantly increased N2 emission from pasture and wetland soils treated with urine, urea and KNO3 at 200 kg N ha-1 rate (Zaman et al., 2007 & 2008c). More recently in a field experiment, a similar trend of enhanced N2 after raising soil pH to 7 was observed in pasture soils treated with urea/urine (Zaman & Nguyen, 2010). This idea is further supported by our studies on cattle pasture soil (Hynšt et al., 2007). At the site with the greatest animal impact, the ratio of N2 to N2O produced during denitrifying enzyme activity (DEA) measurements was five-fold higher, and the pH was 2 units higher, compared to the site with the least animal impact, which indicated that soil conditions were favourable for production of N2 rather than N2O in the area where excretal returns and treading was intense.

Types of chemical N fertilizers are also likely to regulate N2O:N2 ratios, as NH4 + based fertilizers (i.e. ammonium sulphate, ammonium nitrate, and mono-ammonium phosphate) are reported to lower soil pH after their application (Thornton et al., 1996; Mulvaney et al., 1997; Nobre, 2001; Cai et al., 2002). For example, Mulvaney et al. (1997) have reported that ammonium-based fertilizers with soil acidifying effects produce a higher N2O:N2 ratio compared to alkaline forming fertilizers (anhydrous ammonia, urea or di-ammonium phosphate). Most researchers attribute high N2O and low N2 emissions in acidic conditions to the suppression of nitrous oxide-reductase at low soil pH (inhibition at soil pH 4.5) (Kostina et al., 1996; Daum & Schenk 1998; Flessa et al., 1998; Stevens and Laughlin, 1998; Zaman et al., 2007). It is also likely that all denitrifying enzymes are susceptible at low soil pH and produce N2O from intermediate products (Nagele & Conrad, 1990). However, the extensive review conducted by Šimek and Cooper (2002) reported that the lower rates of N2 and high N2O:N2 ratio at low soil pH could be due to lower amounts of soil organic C and mineral N available to the denitrifying population under acid conditions rather than a direct effect of low pH on denitrification enzymes. Regardless of the biochemical reasons for changes in soil pH on N2 emission, raising soil pH through application of soil amendment like lime appears to offer a mechanism for mitigation of N2O (Šimek et al., 2002; Zaman et al., 2007, 2008b, Zaman & Nguyen, 2010).

#### **5.4 Soil aeration and water status**

18 Greenhouse Gases – Emission, Measurement and Management

There are reports that the N2O:N2 ratios are lower in the rhizosphere, which provides more available organic C in the form of root exudates and root debris, and is characterized by low

Depending on the management practices, agricultural soils can act as a source or sink for atmospheric CO2. Improved land management practices in croplands and grasslands can store up to 1 Gt C in the soil on an annual basis (IPCC, 2000). It is therefore possible to store between 100 to 1000 kg SOC ha-1 year-1 depending on the climate, soil and vegetation types, and site-specific soil management practices. Improved land management practices like zero or minimum tillage, retention of crop residues via crop rotation and mulching, application of FDE, organic residues and manure, following crop rotation especially with N fixing crops and avoiding burning crop residues after harvest may offer potential mitigation tool to sequester more C in the soils to offset the increase in atmospheric CO2 as well as to improve soil fertility, soil structure, aggregate stability, pore size geometry and distribution, water and nutrients holding capacity and soil quality (increased microbial and enzymatic activities). Such improvement in soil physical and chemical fertility and health will minimize conducive conditions like anaerobicity and soil compaction which stimulate denitrification. Increased soil C may also help to shift the balance between harmful N2O and non-greenhouse gas N2 during denitrification as the activity of nitrous oxide reductase

Soil pH is among the key regulators of the microbiological processes that affect N2O and N2 production and their ratios. Nitrification activity is generally higher with higher soil pH (> 6) (Bremner & Blackmer, 1981; Bramley & White, 1989). The critical soil pH threshold for nitrification is 5; however, nitrification can occur even at a soil pH of 4.5 due to acid-adapted nitrifier strains (Bouwman, 1990). Denitrification has been reported to occur over a wide range of soil pH values (5 to 8) (Weier & Gilliam, 1986; Ramos, 1996; Flessa et al, 1998); however, laboratory experiments with artificially adjusted soil pH suggest, that under optimized

pHs below 4 or above 10 (Šimek & Hopkins, 1999, Šimek et al., 2002). Numerous laboratory and field studies have shown that soil pH affects N2O and N2 and the ratio of these gases (e.g. Weier & Gilliam, 1986; Stevens & Laughlin, 1998). Under controlled environment experiments, we found that raising soil pH to 7 through lime application significantly increased N2 emission from pasture and wetland soils treated with urine, urea and KNO3 at 200 kg N ha-1 rate (Zaman et al., 2007 & 2008c). More recently in a field experiment, a similar trend of enhanced N2 after raising soil pH to 7 was observed in pasture soils treated with urea/urine (Zaman & Nguyen, 2010). This idea is further supported by our studies on cattle pasture soil (Hynšt et al., 2007). At the site with the greatest animal impact, the ratio of N2 to N2O produced during denitrifying enzyme activity (DEA) measurements was five-fold higher, and the pH was 2 units higher, compared to the site with the least animal impact, which indicated that soil conditions were favourable for production of N2 rather than N2O in the area where excretal

Types of chemical N fertilizers are also likely to regulate N2O:N2 ratios, as NH4

fertilizers (i.e. ammonium sulphate, ammonium nitrate, and mono-ammonium phosphate)

partial pressure of O2 (due to O2 consumption by plant roots) (Casella et al., 1984).

concentration and enzyme status (Swerts et al., 1996).

and glucose amendments), denitrification can proceed even at

+ based


interaction of C availability, NO3

enzymes is stimulated by available soil C.

conditions (very low pO2, NO3

returns and treading was intense.


**5.3 Soil pH** 

Soil aeration, namely O2 concentration and gas exchange between soil and atmosphere, affects all microbial N transformation processes including nitrification, denitrification, and DNRA. Control of the denitrification enzymes, especially nitrous oxide reductase, represents the key mitigation option for the rate of N2O production and can therefore shift the balance between harmful N2O and non-greenhouse N2 production across the agricultural landscape (Smith & Tiedje, 1979; Mosier et al., 1986; Robertson & Tiedje, 1987; Henrich & Haselwandter, 1997; Bollmann & Conrad, 1998; Mosier et al., 2002).). In soil, O2 concentration changes with soil moisture content and organic matter decomposition by soil microorganisms. After rainfall or applying irrigation water, soils become temporarily anaerobic; the extent and duration of anaerobiosis differs with soil types (drainage class). Fine-textured soils with a higher clay content are reported to remain anaerobic for a longer period of time at low WFPS than coarse-textured soils because of the greater number of micro pores in the former (Barton et al., 1999). Therefore fine-textured soils with poorly drained conditions are likely to emit more N2O for a longer period than those of coarsetextured soils with well-drained conditions (Groffman & Tiedje, 1989; Aulakh et al., 1991; Clayton et al., 1997; Dobbie & Smith, 2001; Saggar et al., 2004a). At higher O2, partial pressure (>0.5 vol. %), nitrification is expected to proceed, provided there is sufficient water for optimum activity of nitrifiers (Linn & Doran, 1984; Bollmann & Conrad, 1998); if the soil WFPS increases (and pO2 decreases), the rate of N2O production and the proportion of N2O to NO3- produced also increases (Smith et al., 2003). Under such specific conditions at WFPS>60%, nitrification is considered to be the predominant source of N2O as opposed to denitrification or DNRA (Inubushi et al., 1996). Although DNRA is understood to be an anaerobic process, information about the critical levels of WFPS or O2 for DNRA is lacking in the literature. Denitrification becomes a major source of N2O and N2 production at lower O2 partial pressure (<0.5 vol. %) and higher WFPS (>60%) (Davidson, 1993; Scholefield et al., 1997; Bronson & Fillery, 1998; Khalil et al., 2002). In such scenarios, more aerobic soils are likely to produce mainly N2O because denitrification reductases (Eq. 3.2.1) especially nitrous oxide reductase is reported to be sensitive to soil O2 level, while anaerobic soils and

Emissions of Nitrous Oxide (N2O) and Di-Nitrogen (N2) from

chapter and the unknown reviewers for their positive comments.

*Journal of Environmental Quality,* 31, pp. 1017-1024.


offers potential tools for mitigation of N2O.

lower mineral NO3

**8. References** 

**7. Acknowledgments** 

264, pp.801-802.

pp. 157-163.

1093.

pp. 265-271.

*Journal,* 54, pp. 769-774.

the Agricultural Landscapes, Sources, Sinks, and Factors Affecting N2O and N2 Ratios 21

organic C, soil pH, and soil aeration status. N2 production dominates over that of N2O at

lowering O2 partial pressure or increasing WFPS; above 70%. Manipulation of these factors

The senior author acknowledges the financial assistance for page charges by Ballance Agri-Nutrients Ltd. We also acknowledge Sharon Long for her assistance in proof reading this

Addy, K. Kellogg, D.O. Gold, A.J. Groffman, P.M. Ferendo, G. & Sawyer, C. (2002). In situ

Alef, K. Kleiner, D. (1986). Arginine ammonification, a simple method to estimate microbial activity potential in soils. *Soil Biology and Biochemistry,* 18, pp. 233-235. Arah, J.R.M. Smith, K.A. Crichton, L.J. & Li, H.S. (1991). Nitrous oxide production and denitrification in Scottish soils. *Journal of Soil Science,* 42, pp. 351-367. Armstrong, W. (1964). Oxygen diffusion from the roots of some British bog plants. *Nature,*

Aulakh, M.S. Doran, J.W. & Mosier, A.R. (1991). In-field evaluation of four methods for measuring denitrification. *Soil Science Society of America Journal,* 55, pp. 1332-1338. Aulakh, M.S. Khera, T.S. Doran, J.W. & Bronson, K.F. (2001). Denitrification, N2O and CO2

Azam, F. Malik, K.A. & Hussain, F. (1986). Microbial biomass and mineralization-

Barak, P. Molina, J.A.E. Hadas, A. & Clapp, C.E. (1990). Mineralization of amino acids and

Barraclough, D. (1997). The direct or MIT route for nitrogen immobilization, An 15N mirror

Barton, L. McLay, C.D.A. Schipper, L.A. & Smith, C.T. (1999). Annual denitrfication rates in

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legume green manure. *Biology and Fertility of Soils,* 34, pp. 375-389.

fluxes in rice, wheat cropping system as affected by crop residues, fertilizer N and

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J.L. (2009). Hippuric acid and benzoic acid inhibition of urine derived N2O

influenced by compaction. *Proceedings of the New Zealand Grassland Association,* 65,

pushpull method to determine ground water denitrification in riparian zones.

content, increasing organic C contents, increasing soil pH (above 6.5),

sediments will generate both N2O and N2. A number of studies have reported higher amounts of N2 than N2O at lower O2 partial pressure and WFPS above 70% (Eriksen & Hartwig, 1993; Dendooven et al., 1999; Kwong et al 1999, Khalil et al., 2002). Aulakh et al. (2001) reported that gaseous N losses as N2O after application of urea (120 kg ha-1) to flooded rice were 8 to 10 times higher than those of upland wheat because of the anaerobic conditions in the former. Recently we found that riparian wetland soils treated with NO3--N (200 kg N ha-1 rate) emitted 4 and 8 times more N2O and N2 respectively than pasture soils during 28-day incubation (Zaman et al., 2008 c). However, the relative production of N2O and N2 in anaerobic or aerobic soil conditions is not that simple since O2 level or WFPS is only one of the many known soil and management factors which affect this relationship (Fillery, 1983; Scholefield et al., 1997; Stevens & Laughlin, 1998; Zaman et al., 2008b). In their comprehensive review on emissions of N2O and NO from fertilized fields published, Bouwman et al. (2002) concluded that restricted drainage and fine texture favors N2O emissions. Thus our current understanding of the processes of N2O and N2 production in anaerobic and aerobic soil conditions is limited. At this stage we can only suggest that improving soil drainage conditions and avoiding soil compaction through use of the heavy farm machinery and grazing animals (pugging) in wet soil conditions (especially in winter) could help to maintain aerobicity in soils, which in turn may reduce N2O emission rates through nitrification, denitrification and DNRA (Bhandral et al., 2003, 2007b. Luo et al., 2008b).

Apart from the above mentioned factors, temperature is also known to affect N2O production and the N2O:N2 ratio (Cho et al., 1997; Daum et al (1997, Muller et al., 2002). However, controlling soil temperature is mostly beyond the ability of farmers. Manipulation of the interaction between mineral N supply (NH4+ and NO3-), organic C, soil aeration and pH offers the best hope for minimizing N2O emission from soils. Similarly the export of N via surface and sub-surface runoff from upland to water bodies can be minimized by using riparian zones (both natural and constructed) along river and stream banks. Since denitrification is considered to be the major NO3- removal process in wetland, proper management of wetlands include, regular application of lime to keep the pH above 6.5, sequestering C to build C reserves, and exclusion of grazing animals to minimize N inputs are essential. All these management practices are known to stimulate the activity of nitrous oxide reductase, which will help to result in emissions of more N2 than N2O as discussed above.

## **6. Conclusions**

Nitrogen is the most dynamic plant, microbial and animal nutrient which affects the diversity, dynamics, and functioning of many terrestrial, freshwater and marine ecosystems. Gaseous N losses in the form of N2O are undesirable because N2O is an important greenhouse gas and is also involved in the depletion of stratospheric ozone. Nitrification, denitrification and DNRA are the main microbial processes for N2O production across the agricultural landscape which can sometimes operate concurrently in a given soil system. N losses as N2O across the agricultural landscape are extremely variable and range from about 1% to more than 20 % of the applied N. Such losses are generally higher from wetland soils than those from pasture or arable soils. The critical soil and management factors affecting the rates of N2O and N2 production and their ratios are concentration of mineral N, soil organic C, soil pH, and soil aeration status. N2 production dominates over that of N2O at lower mineral NO3 content, increasing organic C contents, increasing soil pH (above 6.5), lowering O2 partial pressure or increasing WFPS; above 70%. Manipulation of these factors offers potential tools for mitigation of N2O.

## **7. Acknowledgments**

The senior author acknowledges the financial assistance for page charges by Ballance Agri-Nutrients Ltd. We also acknowledge Sharon Long for her assistance in proof reading this chapter and the unknown reviewers for their positive comments.

## **8. References**

20 Greenhouse Gases – Emission, Measurement and Management

sediments will generate both N2O and N2. A number of studies have reported higher amounts of N2 than N2O at lower O2 partial pressure and WFPS above 70% (Eriksen & Hartwig, 1993; Dendooven et al., 1999; Kwong et al 1999, Khalil et al., 2002). Aulakh et al. (2001) reported that gaseous N losses as N2O after application of urea (120 kg ha-1) to flooded rice were 8 to 10 times higher than those of upland wheat because of the anaerobic conditions in the former. Recently we found that riparian wetland soils treated with NO3--N (200 kg N ha-1 rate) emitted 4 and 8 times more N2O and N2 respectively than pasture soils during 28-day incubation (Zaman et al., 2008 c). However, the relative production of N2O and N2 in anaerobic or aerobic soil conditions is not that simple since O2 level or WFPS is only one of the many known soil and management factors which affect this relationship (Fillery, 1983; Scholefield et al., 1997; Stevens & Laughlin, 1998; Zaman et al., 2008b). In their comprehensive review on emissions of N2O and NO from fertilized fields published, Bouwman et al. (2002) concluded that restricted drainage and fine texture favors N2O emissions. Thus our current understanding of the processes of N2O and N2 production in anaerobic and aerobic soil conditions is limited. At this stage we can only suggest that improving soil drainage conditions and avoiding soil compaction through use of the heavy farm machinery and grazing animals (pugging) in wet soil conditions (especially in winter) could help to maintain aerobicity in soils, which in turn may reduce N2O emission rates through nitrification, denitrification and DNRA (Bhandral et al., 2003, 2007b. Luo et al.,

Apart from the above mentioned factors, temperature is also known to affect N2O production and the N2O:N2 ratio (Cho et al., 1997; Daum et al (1997, Muller et al., 2002). However, controlling soil temperature is mostly beyond the ability of farmers. Manipulation of the interaction between mineral N supply (NH4+ and NO3-), organic C, soil aeration and pH offers the best hope for minimizing N2O emission from soils. Similarly the export of N via surface and sub-surface runoff from upland to water bodies can be minimized by using riparian zones (both natural and constructed) along river and stream banks. Since denitrification is considered to be the major NO3- removal process in wetland, proper management of wetlands include, regular application of lime to keep the pH above 6.5, sequestering C to build C reserves, and exclusion of grazing animals to minimize N inputs are essential. All these management practices are known to stimulate the activity of nitrous oxide reductase, which will help to result in emissions of more N2 than N2O as discussed

Nitrogen is the most dynamic plant, microbial and animal nutrient which affects the diversity, dynamics, and functioning of many terrestrial, freshwater and marine ecosystems. Gaseous N losses in the form of N2O are undesirable because N2O is an important greenhouse gas and is also involved in the depletion of stratospheric ozone. Nitrification, denitrification and DNRA are the main microbial processes for N2O production across the agricultural landscape which can sometimes operate concurrently in a given soil system. N losses as N2O across the agricultural landscape are extremely variable and range from about 1% to more than 20 % of the applied N. Such losses are generally higher from wetland soils than those from pasture or arable soils. The critical soil and management factors affecting the rates of N2O and N2 production and their ratios are concentration of mineral N, soil

2008b).

above.

**6. Conclusions** 


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**The Blend Ethanol/Gasoline and**

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The blended ethanol gasoline has major effects on reducing emissions of greenhouse gases, while the engine design, as well as the operation conditions, and the presence of a catalyst in

The preliminary combustion processes are explained by reactions mechanisms and chemical kinetics, and are important to show how these gases are formed and to determine the

The engine has a decisive role in the efficient burning; in this case, the physical and chemical properties of the fuel must be specified in accordance with the requirements of your project. On the other hand, in Otto cycle engines, adding ethanol to gasoline, it reduces emissions of greenhouse gases like carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O), and pollutants like carbon monoxide (CO), volatile organic compounds (VOC). Though, there is

The purpose of this study is to compare the emissions of pollutant and greenhouse gases, when using the blend gasoline/ethanol or hydrated ethanol; understanding the differences between the combustion of two types of vehicles, and reviewing mechanisms that allow the

as intake, compression (combustion), expansion and exhaustion.

due to expansion in the exhaust and then to the environment.

the exhaust, decrease pollutants and greenhouse gases.

an increase in nitrogen oxides (NOx), and aldehydes (RCHO).

increments and reduction of pollutant and the greenhouse gases.

**1. Introduction** 

exhaust.

emission levels.

**Emission of Gases** 

Antonio Carlos Santos *Empresa de Pesquisa Energética* 

