**Methane Emissions from Rice Production in the United States — A Review of Controlling Factors and Summary of Research**

Alden D. Smartt, Kristofor R. Brye and Richard J. Norman

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/62025

#### **Abstract**

Flooded rice (*Oryza sativa* L.) cultivation has been identified as one of the leading global agricultural sources of anthropogenic methane (CH4) emissions. Furthermore, it has been estimated that global rice production is responsible for 11% of total anthropogenic CH4 emissions. Considering that CH4 has a global warming potential that is approximately 25 times more potent, on a mass basis, than carbon dioxide (CO2) and rice production is globally extensive and concentrated in several mid-southern and southern states and Cal‐ ifornia, the purpose of this review is two-fold: (i) discuss the factors known to control CH4 production in the soil and transport to the atmosphere from rice cultivation and (ii) summarize the historic and recent research conducted on CH4 emissions from rice pro‐ duction in the temperate United States. Though some knowledge has been gained, there is much more that still needs to be learned and understood regarding CH4 emissions from rice production in the United States, its contribution to climate change, and poten‐ tial mitigation strategies. Extending the current knowledge base surrounding CH4 emis‐ sions from rice cultivation will help regulatory bodies, such as the Environmental Protection Agency, refine greenhouse gas emissions factors to combat the potential nega‐ tive effects of climate change.

**Keywords:** Methane, emissions, rice production, agriculture, soil texture

#### **1. Introduction**

Methane (CH4) is a known and potent greenhouse gas that is produced by anaerobic *Archaea* under anoxic conditions. Agricultural activities have been recognized as contributing an estimated 50% to global anthropogenic CH4 emissions [1], while an estimated 31% of anthro‐ pogenic CH4 emissions have been attributed to agricultural activities in the United States (US)

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[2]. Due to the anaerobic conditions that form in saturated soils, which is a prerequisite for CH4 production, flooded rice (*Oryza sativa* L.) cultivation has been specifically identified as one of the leading global agricultural sources of anthropogenic CH4 emissions, accounting for approximately 22% of the total global agriculturally related CH4 emissions [3]. Furthermore, it has been estimated that global rice production is responsible for 11% of total anthropogenic CH4 emissions [1,3].

While numerous factors have been determined to impact CH4 emissions from rice cultivation, due to a general lack of field data, the United States Environmental Protection Agency (USEPA) currently uses a single emissions factor for all non-California-grown, primary rice crops [4]. Therefore, the purpose of this review is two-fold: (i) discuss the factors known to control CH4 production in the soil and transport to the atmosphere from rice cultivation and (ii) summarize the historic and recent research conducted on CH4 emissions from rice production in the temperate United States.

#### **2. The greenhouse effect**

The greenhouse effect is a mechanism by which certain gases such as carbon dioxide (CO2), CH4, nitrous oxide (N2O), and water (H2O) vapor absorb and release infrared radiation, interfering with the ability of solar radiation to leave Earth's atmosphere. The absorption of thermal radiation by H2O and CO2 was discovered through laboratory experiments in 1859 [5]. However, other gases including CH4 and N2O were not recognized as greenhouse gases until the 1970s [6].

Global warming potential (GWP) is a metric that allows the warming impact of various greenhouse gases to be quantitatively compared on the same scale. The assignment of GWP values to gases requires knowledge of the contribution to global warming of gas emissions over time based on the amount of radiation per mass that the gas can absorb and emit as well as the atmospheric lifetime of the gas. Global warming potentials are assigned relative to that of CO2, thus the 100-yr GWP of CO2, CH4, and N2O are 1, 25, and 298, respectively [7]. For example, 1 kg of CH4 released to the atmosphere is equivalent to 25 kg of CO2 being released. Global warming potentials allow greenhouse gas emissions to be reported as CO2 equivalents in order to compare warming effects of various gases on a single scale.

The current climate change problem is not a result of the greenhouse effect itself, but rather from an increasing greenhouse effect resulting from anthropogenic activities that have increased atmospheric concentrations of greenhouse gases. Prior to 1750, the atmospheric CO2 mixing ratio was about 280 parts per million (ppm) [8]. Since the beginning of the industrial era, atmospheric CO2 has risen drastically to 379 ppm in 2005 [7] and 395 ppm as of April 2013 [9]. Between 1750 and 2005, atmospheric CH4 increased from about 700 parts per billion (ppb) to 1,774 ppb [7]. Nitrous oxide was more variable ranging from 180 to 260 ppb prior to 1750, but has similarly increased to a mixing ratio of 319 ppb in 2005 [7]. While atmospheric N2O and CO2 concentrations have increased steadily over the past several decades, the growth rate (i.e., concentration increase) of atmospheric CH4 seems to be declin‐ ing. The growth rate of atmospheric CH4 has decreased from highs of about 1% per year in the 1970s and 1980s to nearly zero between 1999 and 2005. However, the decreasing growth rate is poorly understood [7].

#### **3. Greenhouse gas emissions**

[2]. Due to the anaerobic conditions that form in saturated soils, which is a prerequisite for CH4 production, flooded rice (*Oryza sativa* L.) cultivation has been specifically identified as one of the leading global agricultural sources of anthropogenic CH4 emissions, accounting for approximately 22% of the total global agriculturally related CH4 emissions [3]. Furthermore, it has been estimated that global rice production is responsible for 11% of total anthropogenic

While numerous factors have been determined to impact CH4 emissions from rice cultivation, due to a general lack of field data, the United States Environmental Protection Agency (USEPA) currently uses a single emissions factor for all non-California-grown, primary rice crops [4]. Therefore, the purpose of this review is two-fold: (i) discuss the factors known to control CH4 production in the soil and transport to the atmosphere from rice cultivation and (ii) summarize the historic and recent research conducted on CH4 emissions from rice production

The greenhouse effect is a mechanism by which certain gases such as carbon dioxide (CO2), CH4, nitrous oxide (N2O), and water (H2O) vapor absorb and release infrared radiation, interfering with the ability of solar radiation to leave Earth's atmosphere. The absorption of thermal radiation by H2O and CO2 was discovered through laboratory experiments in 1859 [5]. However, other gases including CH4 and N2O were not recognized as greenhouse gases until

Global warming potential (GWP) is a metric that allows the warming impact of various greenhouse gases to be quantitatively compared on the same scale. The assignment of GWP values to gases requires knowledge of the contribution to global warming of gas emissions over time based on the amount of radiation per mass that the gas can absorb and emit as well as the atmospheric lifetime of the gas. Global warming potentials are assigned relative to that of CO2, thus the 100-yr GWP of CO2, CH4, and N2O are 1, 25, and 298, respectively [7]. For example, 1 kg of CH4 released to the atmosphere is equivalent to 25 kg of CO2 being released. Global warming potentials allow greenhouse gas emissions to be reported as CO2 equivalents

The current climate change problem is not a result of the greenhouse effect itself, but rather from an increasing greenhouse effect resulting from anthropogenic activities that have increased atmospheric concentrations of greenhouse gases. Prior to 1750, the atmospheric CO2 mixing ratio was about 280 parts per million (ppm) [8]. Since the beginning of the industrial era, atmospheric CO2 has risen drastically to 379 ppm in 2005 [7] and 395 ppm as of April 2013 [9]. Between 1750 and 2005, atmospheric CH4 increased from about 700 parts per billion (ppb) to 1,774 ppb [7]. Nitrous oxide was more variable ranging from 180 to 260 ppb prior to 1750, but has similarly increased to a mixing ratio of 319 ppb in 2005 [7]. While atmospheric N2O and CO2 concentrations have increased steadily over the past several decades, the growth rate (i.e., concentration increase) of atmospheric CH4 seems to be declin‐

in order to compare warming effects of various gases on a single scale.

CH4 emissions [1,3].

180 Greenhouse Gases

in the temperate United States.

**2. The greenhouse effect**

the 1970s [6].

Globally, CO2 accounted for about 76% of greenhouse gas emissions in 2004, with around 75% of CO2 emissions resulting from fossil fuel use and much of the remainder from deforestation and biomass decomposition [10]. Methane and N2O accounted for 14 and 8%, respectively, of the estimated global greenhouse gas emissions in 2004. Major CH4 sources include agricultural activities, waste management, and energy use, while N2O emissions are primarily a result of agricultural activities, such as fertilizer use and soil management [10]. In the US in 2013, an estimated 82% of the total greenhouse gas emissions were CO2, 10% were CH4, and 5% were N2O [2]. Major sources of greenhouse gas emissions are generally the same in the US as the global sources mentioned above. The major global sectors responsible for greenhouse gas emissions are energy supply (26%), industry (19%), forestry (17%), agriculture (14%), and transportation (13%) [10]. In comparison, the major US sectors responsible for greenhouse gas emissions are energy supply (31%), transportation (27%), industry (21%), commercial and residential (12%), and agriculture (9%) [2].

Although agricultural activities do not dominate total greenhouse gas emissions, agriculture contributes an estimated 50 and 60% of global anthropogenic emissions of CH4 and N2O, respectively [1]. Agriculture in the US is responsible for an estimated 36% of anthropogenic CH4 emissions and 79% of anthropogenic N2O emissions [2]. Enteric fermentation, rice cultivation, and manure management contribute an estimated 64, 22, and 8%, respectively, to global anthropogenic agricultural CH4 emissions, while agricultural N2O emissions are dominated by agricultural soil management (80%) [3]. In comparison, enteric fermentation, rice cultivation, and manure management contribute to 70, 4, and 26% of US anthropogenic agricultural CH4 emissions [2]. Although rice cultivation makes up a small portion of CH4 emissions in the US, globally rice cultivation accounts for approximately 11% of total anthro‐ pogenic CH4 emissions.

Methane emissions from US rice cultivation were estimated to be 8.3 Tg CO2 equivalents in 2013, a reduction from 9.3 Tg CO2 equivalents in 2012 due to a decline in rice production area [2]. Arkansas was responsible for 36% of the estimated CH4 emissions from rice cultivation, although Arkansas accounted for 43% of the total US rice production in 2013. Louisiana was the next leading contributor to CH4 emissions accounting for 27% of 2013 emissions, while harvesting 16% of 2013 production [2,11]. Louisiana and Texas CH4 emissions are large relative to their production areas due to extensive ratoon cropping in 2013, which occurred on an estimated 38 and 68%, respectively, of the production area in those states [2]. A ratoon crop is a second crop that is managed and produced after the first or primary crop is harvested. California, Mississippi, and Missouri, none of which reported any ratoon cropping, contrib‐ uted 14, 3.6, and 4.5%, respectively, to the estimated 2013 CH4 emissions from US rice culti‐ vation [2].

The USEPA periodically publishes CH4 emissions factors based on research data. Separate emission factors of 178 kg CH4-C ha–1 season–1 and 585 kg CH4-C ha–1 season–1 were used in the inventory estimates for non-California-grown, primary rice cropping and ratooned cropping areas, respectively, as is consistent with the Intergovernmental Panel on Climate Change [3], which recommends calculating separate emissions factors for as many different factors and cultural practices as is possible. Emissions factors for California rice production are 200 and 100 kg CH4-C ha–1 season–1 for winter-flooded and non-winter-flooded rice, respectively [2]. While it is known that factors such as water management, soil properties, rice cultivar, fertilizer management, and residue management have strong impacts on CH4 emissions from rice cultivation, data available from US studies limit the further disaggregation of these factors [2]. The non-California-grown, primary crop emissions factor is based on US studies with emissions ranging from 46 to 375 kg CH4-C ha–1 season–1 [13–20] and the ratoon crop factor is based on studies conducted in Louisiana with emissions ranging from 361 to 1118 kg CH4-C ha–1 season–1 [21,22]. The California-specific emissions factors include studies with emissions ranging from 47 to 166 kg CH4-C ha–1 season–1 for the non-winter-flooded and from 98 to 277 kg CH4-C ha–1 season–1 for the winter-flooded rice [23,24].

#### **4. Rice production**

Rice is a semi-aquatic, cereal grain that makes up about 21% of total global grain production [25]. The importance of rice is further exemplified by the fact that rice is a staple food crop for about half of the global population, with direct human consumption accounting for 85% of rice production compared to 72% of wheat (*Triticum aestivum* L.) and 19% of maize (*Zea mays* L.) production [26,27]. In Southeast Asia, 60% of human food intake is provided by rice as well as 35% of food intake in both East Asia and South Asia [26]. Rice has the ability to support more people per unit of land area than wheat or maize because rice produces, on an average yield basis, more food energy and protein per hectare than wheat or maize [28]. Therefore, any potential negative environmental consequences associated with rice production have to be taken seriously.

#### **4.1. Rice production extent**

Common rice (*Oryza sativa*) is commercially produced in 112 countries worldwide, spanning latitudes from 53°N along the Amur River at the China–Russia border to 35°S in central Argentina [26]. In 2012, more than 158 million ha globally were planted to rice, with average yields of 4.4 Mg ha–1 for a total global production of 470 Tg of rice. Comparatively, nearly 216 million ha were planted in wheat in 2012, with average yields of 3.0 Mg ha–1 for a total of 656 Tg of global wheat production. More than 174 million ha were planted in maize in 2012, with an average yield of 4.9 Mg ha–1 and a total global production of 857 Tg of maize [25]. Global rice production peaked in 1994 at 534 Tg of rice, with Asia being responsible for 90% of that production [29]. The majority of global rice production occurs in east, south, and southeast Asia, which together accounted for 90% of global production in 2012. Substantial production also occurs in South America (Brazil and Peru), Sub-Saharan Africa (Nigeria and Madagascar), Europe (Italy and Spain), Egypt, and the US [25].

China and India currently dominate global rice production accounting for 30 and 22%, respectively, of the total global production in 2012. The third-, fourth-, and fifth-ranked global producers in 2012 were Indonesia (8%), Bangladesh (7%), and Vietnam (6%). The remaining top 10 producers, in order, were Thailand, the Philippines, Burma, Brazil, and Japan, followed by the eleventh-ranked US, which accounted for 1.3% of global production [25]. The US, however, plays a larger role in global exports contributing 9% of 2012 global exports and ranking fifth after Thailand (21%), India (20%), Vietnam (20%), and Pakistan (10%). Global exports in 2012 were estimated to be 8% of total production, while the US exported 55% of 2012 production [30]. Global rice yields in 2012 were estimated to be 4.4 Mg ha–1 compared to 8.3 Mg ha–1 in the US, which was second only to Egypt (8.8 Mg ha–1) among the major rice-growing countries. The two top rice-producing countries, China and India, had estimated yields of 6.7 and 3.6 Mg ha–1, respectively [25].

Nearly 1.1 million ha of rice were planted in the US in 2012, yielding an average of 8.3 Mg ha–1 for a total production of 9.0 Tg of rice prior to milling, compared to 23 million ha planted with an average yield of 3.1 Mg ha–1 for a total of 62 Tg of wheat production, and over 39 million ha of planted maize with average yields of 7.7 Mg ha–1 for a total production of 274 Tg [11]. The four major regions that produce rice in the US are the Arkansas Grand Prairie, the Mississippi Delta, which is made up of portions of Arkansas, Missouri, Mississippi, and Louisiana, the Gulf Coast (Texas and southwest Louisiana), and California's Sacramento Valley. Most US states produce primarily long-grain cultivars, while much of the mediumgrain rice and nearly all of the short-grain rice is produced in California [11]. Although Oklahoma and Florida are often included as rice-producing states, the six previously men‐ tioned states have made up essentially all of US production in recent years [11]. Arkansas is the leading state in both area of cultivation and total production, contributing 48% of total US rice production in 2012, followed by 23% of production by California and 13% of production by Louisiana [11]. Arkansas rice production takes place in the eastern portion of the state with the top five rice-producing counties in 2012 being Poinsett, Lawrence, Arkansas, Greene, and Cross, which made up 35% of the state's production area [31].

#### **4.2. Global rice production practices**

The USEPA periodically publishes CH4 emissions factors based on research data. Separate emission factors of 178 kg CH4-C ha–1 season–1 and 585 kg CH4-C ha–1 season–1 were used in the inventory estimates for non-California-grown, primary rice cropping and ratooned cropping areas, respectively, as is consistent with the Intergovernmental Panel on Climate Change [3], which recommends calculating separate emissions factors for as many different factors and cultural practices as is possible. Emissions factors for California rice production are 200 and 100 kg CH4-C ha–1 season–1 for winter-flooded and non-winter-flooded rice, respectively [2]. While it is known that factors such as water management, soil properties, rice cultivar, fertilizer management, and residue management have strong impacts on CH4 emissions from rice cultivation, data available from US studies limit the further disaggregation of these factors [2]. The non-California-grown, primary crop emissions factor is based on US studies with emissions ranging from 46 to 375 kg CH4-C ha–1 season–1 [13–20] and the ratoon crop factor is based on studies conducted in Louisiana with emissions ranging from 361 to 1118 kg CH4-C ha–1 season–1 [21,22]. The California-specific emissions factors include studies with emissions ranging from 47 to 166 kg CH4-C ha–1 season–1 for the non-winter-flooded and from 98 to 277

Rice is a semi-aquatic, cereal grain that makes up about 21% of total global grain production [25]. The importance of rice is further exemplified by the fact that rice is a staple food crop for about half of the global population, with direct human consumption accounting for 85% of rice production compared to 72% of wheat (*Triticum aestivum* L.) and 19% of maize (*Zea mays* L.) production [26,27]. In Southeast Asia, 60% of human food intake is provided by rice as well as 35% of food intake in both East Asia and South Asia [26]. Rice has the ability to support more people per unit of land area than wheat or maize because rice produces, on an average yield basis, more food energy and protein per hectare than wheat or maize [28]. Therefore, any potential negative environmental consequences associated with rice production have to be

Common rice (*Oryza sativa*) is commercially produced in 112 countries worldwide, spanning latitudes from 53°N along the Amur River at the China–Russia border to 35°S in central Argentina [26]. In 2012, more than 158 million ha globally were planted to rice, with average yields of 4.4 Mg ha–1 for a total global production of 470 Tg of rice. Comparatively, nearly 216 million ha were planted in wheat in 2012, with average yields of 3.0 Mg ha–1 for a total of 656 Tg of global wheat production. More than 174 million ha were planted in maize in 2012, with an average yield of 4.9 Mg ha–1 and a total global production of 857 Tg of maize [25]. Global rice production peaked in 1994 at 534 Tg of rice, with Asia being responsible for 90% of that production [29]. The majority of global rice production occurs in east, south, and southeast Asia, which together accounted for 90% of global production in 2012. Substantial production

kg CH4-C ha–1 season–1 for the winter-flooded rice [23,24].

**4. Rice production**

182 Greenhouse Gases

taken seriously.

**4.1. Rice production extent**

Rice production practices vary globally based on economic, cultural, and climatic factors, each of which show temporal and spatial variability throughout the rice-growing countries. A simple classification or characterization of rice production systems is nearly impossible on a global scale due to the variability of factors that influence production. Classifications of rice production techniques are commonly based upon flood presence (e.g., upland or lowland), water source (e.g., irrigated or rainfed), and stand establishment technique (e.g., transplanting, direct-seeding, or water-seeding) with many combinations and variations of these techniques occurring throughout the globe [32]. In one of the most recent classification attempts, Chang [33] classified global rice production into five major agroecosystems: (i) irrigated wetland, which made up 53% of global rice production area and had the greatest yield potential at 3 to 5 Mg ha–1, (ii) rainfed wetland, making up 26% of global area and yielding 2 to 4 Mg ha–1, (iii) flood-prone or tidal swamps, which made up an insignificant area, (iv) deep water (1–5 m), making up 8% of global area, and (v) dryland, which made up an estimated 13% of global production area with average yield potentials of 1 to 1.5 Mg ha–1.

While a small portion of rice is produced under upland conditions, the majority of rice production requires substantial quantities of water in order to maintain a flood on the semiaquatic crop. In much of the tropical rice-growing area, particularly south and southeast Asia, rainfed rice is the main production system, where most of the production comes from wetseason harvests and the cropping season is determined by rainfall patterns [32]. In temperate production areas, rice production must coincide with suitable temperatures for the crop which, coupled with inadequate rainfall, requires that temperate rice be almost entirely irrigated in order to maintain a flood for the duration of the growing season [32]. The utilization of irrigation in temperate areas allows greater control of environmental factors, which ultimately tends to increase yields, while rainfed systems may suffer from droughts and floods that may substantially damage crops and reduce yields [32].

Direct-seeding and transplanting are common establishment techniques in both irrigated- and rainfed-wetland systems, while direct-seeding is the major practice in dryland and deep-water agroecosystems [33]. While transplanting does occur in irrigated- and direct seeding occurs in rainfed-wetland systems, it is more common for irrigated systems to utilize direct-seeding and for rainfed systems to use transplanting techniques [32]. Transplanting systems involve raising seedlings in a nursery seedbed area at the beginning of the season and transplanting into puddled paddy soils early in the vegetative growth stage. Transplanting is the major estab‐ lishment system for rainfed rice in tropical Asia, with the majority of production in northeast India, Bangladesh, and Thailand relying upon transplanting techniques [32]. Direct-seeding by grain-drilling or broadcasting pre-germinated seeds onto puddled soil is practiced in parts of India, Sri Lanka, Bangladesh, and the Philippines, while drill-seeding into dry soil is the most common practice in the US and other mechanized regions such as Australia [32]. Rice seed may be broadcast onto dry or moist soil by airplane followed by harrowing to cover seeds, but this establishment method requires more seed and stand establishment is often poorer than with drill-seeding [32]. Water-seeding is an establishment technique that originated and is practiced in parts of Asia, where pre-germinated seeds are broadcasted from an airplane into already flooded paddies or fields [32]. The rice-production system, and associated specific production practices, can significantly affect CH4 production and emissions.

#### **4.3. Rice production practices in the US**

Rice production under mechanized US systems requires high temperatures, nearly level land, plentiful water, and soils that inhibit percolation of floodwater, so production is limited to Arkansas, Louisiana, Mississippi, Missouri, Texas, California, and Florida [34]. All US rice is produced using high-input, mechanized production practices, but practices vary somewhat from region to region based on differences in climate, soils, weed proliferation, and other factors that influence production. Essentially, all US rice is irrigated and sources of irrigation water include shallow or deep groundwater, runoff reservoirs, rivers, bayous, and lakes [34]. It is estimated that between 1000 and 2500 m3 ha–1 of water are required to produce a rice crop in the southern US and generally less than one third of that requirement is met by rainfall [35]. Levees, which separate fields into bays, or paddies, and control flood depth (i.e., by use of gates or spills), are commonly constructed on contours that were surveyed on 3 to 6 cm vertical intervals. This creates winding, contour-shaped levees in fields that are not precision-leveled, whereas precision leveling to a uniform grade of 0.2% or less allows the construction of uniformly spaced, straight levees and may reduce the number of levees required [34].

5 Mg ha–1, (ii) rainfed wetland, making up 26% of global area and yielding 2 to 4 Mg ha–1, (iii) flood-prone or tidal swamps, which made up an insignificant area, (iv) deep water (1–5 m), making up 8% of global area, and (v) dryland, which made up an estimated 13% of global

While a small portion of rice is produced under upland conditions, the majority of rice production requires substantial quantities of water in order to maintain a flood on the semiaquatic crop. In much of the tropical rice-growing area, particularly south and southeast Asia, rainfed rice is the main production system, where most of the production comes from wetseason harvests and the cropping season is determined by rainfall patterns [32]. In temperate production areas, rice production must coincide with suitable temperatures for the crop which, coupled with inadequate rainfall, requires that temperate rice be almost entirely irrigated in order to maintain a flood for the duration of the growing season [32]. The utilization of irrigation in temperate areas allows greater control of environmental factors, which ultimately tends to increase yields, while rainfed systems may suffer from droughts and floods that may

Direct-seeding and transplanting are common establishment techniques in both irrigated- and rainfed-wetland systems, while direct-seeding is the major practice in dryland and deep-water agroecosystems [33]. While transplanting does occur in irrigated- and direct seeding occurs in rainfed-wetland systems, it is more common for irrigated systems to utilize direct-seeding and for rainfed systems to use transplanting techniques [32]. Transplanting systems involve raising seedlings in a nursery seedbed area at the beginning of the season and transplanting into puddled paddy soils early in the vegetative growth stage. Transplanting is the major estab‐ lishment system for rainfed rice in tropical Asia, with the majority of production in northeast India, Bangladesh, and Thailand relying upon transplanting techniques [32]. Direct-seeding by grain-drilling or broadcasting pre-germinated seeds onto puddled soil is practiced in parts of India, Sri Lanka, Bangladesh, and the Philippines, while drill-seeding into dry soil is the most common practice in the US and other mechanized regions such as Australia [32]. Rice seed may be broadcast onto dry or moist soil by airplane followed by harrowing to cover seeds, but this establishment method requires more seed and stand establishment is often poorer than with drill-seeding [32]. Water-seeding is an establishment technique that originated and is practiced in parts of Asia, where pre-germinated seeds are broadcasted from an airplane into already flooded paddies or fields [32]. The rice-production system, and associated specific

production practices, can significantly affect CH4 production and emissions.

Rice production under mechanized US systems requires high temperatures, nearly level land, plentiful water, and soils that inhibit percolation of floodwater, so production is limited to Arkansas, Louisiana, Mississippi, Missouri, Texas, California, and Florida [34]. All US rice is produced using high-input, mechanized production practices, but practices vary somewhat from region to region based on differences in climate, soils, weed proliferation, and other factors that influence production. Essentially, all US rice is irrigated and sources of irrigation water include shallow or deep groundwater, runoff reservoirs, rivers, bayous, and lakes [34].

production area with average yield potentials of 1 to 1.5 Mg ha–1.

184 Greenhouse Gases

substantially damage crops and reduce yields [32].

**4.3. Rice production practices in the US**

The two stand establishment techniques utilized in the US are dry-seeding and water-seeding. Dry-seeding techniques, particularly drill-seeding, are predominant in most of the US, while water-seeding techniques are used extensively in California and to a small degree in southwest Louisiana and other regions as a weed control method [34]. A continuously flooded, waterseeding technique is used in California, where pre-germinated seeds are broadcast by airplane into flooded fields and the seedlings grow through a standing flood, while a pinpoint-flood, water-seeding technique is used in Louisiana, where seeds are broadcast into a flooded field that is drained within a few days and then permanently flooded after drying for 3 to 5 days [34,36]. In dry-seeded systems, seed is most often drilled into a well-pulverized, firm, and weed-free seedbed in 15- to 25-cm rows to a depth of 2.5 cm or less. When rice is following a high-residue crop, such as rice, maize, or wheat, it is necessary to till the land in the fall or early spring so that decomposition of the residue does not immobilize nutrients after the subsequent rice crop is planted, whereas rice following soybean (*Glycine max* L.), a crop that produces relatively little residue, may not require as much preparation because crop residues are not as abundant or as persistent compared to that of rice or maize [34,37].

Water management at and shortly after planting varies across US systems, but a permanent flood is established in all systems usually by the four- to five-leaf vegetative growth stage/ beginning tillering (V4-5) [38]. Flush irrigation is used as necessary to promote germination and seedling growth in dry-seeded rice systems prior to establishment of a permanent flood, which typically occurs three to four weeks after emergence (i.e., the V4 to V5 growth stage). Drainage during the season is typically avoided except if a nutrient deficiency, such as zinc, is detected, to aerate the soil in order to treat or prevent disorders, such as straighthead and hydrogen sulfide toxicity, or to apply pesticides. Fields are drained prior to harvest in order to dry the soil enough for operation of harvest equipment [34]. Fields are flooded again within five to seven days after primary-crop harvest in ratoon cropping systems, which are common in southwest Louisiana, Texas, and Florida, and the flood is again maintained until harvest of the ratoon crop [34].

Crop rotations are important in rice, especially where weedy/red rice is problematic and difficult to control during rice cropping seasons. In order to suppress weedy rice, nearly all rice in Louisiana is grown either in a 1:1 rotation with soybean or a 1:1:1 rotation where crawfish (*Procambarus clarkia*) are double-cropped following rice, with soybean produced the following season [34]. In 2012, greater than 70% of Arkansas rice was produced in rotation with soybean, with most of the remaining production in a rice–rice rotation [39]. In California, approximately 70% of rice is produced in a rice–fallow or rice–rice rotation [40].

#### **4.4. Arkansas rice production practices**

Arkansas is the leading rice-producing state, accounting for 40 to 50% of total annual produc‐ tion in the US [11]. Rice production in Arkansas began in 1902 when 0.4 ha were planted in Lonoke County. Production increased over time until 1955 when government quotas limited production to 202,350 ha. The limitation was lifted in 1974 and production increased again, peaking in 1981 at 623,240 ha, again in 1999 with 667,755 harvested hectares, and finally in 2010 with 724,413 ha [31]. In 2012, 518,016 ha rice were harvested in Arkansas [11]. Rice production in Arkansas is highly mechanized with a heavy dependence upon synthetic fertilizers, chemical pest control, and machinery. Planting of rice in Arkansas generally begins the last week of March and extends into early June with floods typically being established by the end of May or early June. Harvesting operations usually begin in mid-August and peak in early- to mid-September [31].

Arkansas rice is produced on a wide variety of soils ranging from sandy to clay soils with the differing textural classes generally requiring different management, especially with regards to tillage practices and nutrient management [39, 41]. Production on sands and sandy loams is minor and has been decreasing from 3.1 and 5.2% of Arkansas area, respectively, in 2007 to 0.7 and 3.7%, respectively, in 2012. Arkansas production on clay and clay-loam soils, however, has increased from under 40 to 48% between 2007 and 2009 but declined to 43% in 2012. Production on silt-loam soils has remained fairly steady at 52% in 2007 and 53% in 2012 [39,42].

Dry-seeding techniques have always dominated in Arkansas. Water-seeding has varied between 2 and 8% of the production area between 2007 and 2012, with an estimated 5% of the 2012 Arkansas rice area being water-seeded [39,42]. Approximately 80% of 2012 Arkansas rice area was drill-seeded, compared to approximately 20% being broadcast-seeded [39]. Conven‐ tional tillage accounted for over half of Arkansas planted-rice area, while stale-seedbed (i.e., tillage and floating, or leveling the field, in the fall or winter) and no-tillage accounted for 35 and 10% of the planted-rice area, respectively, in 2012 [39]. Stale-seedbed and no-tillage are oftentimes utilized on clay soils where conventional tillage produces a cloddy seedbed with poor seed-to-soil contact [41].

While pinpoint, water-seeding techniques do occur in Arkansas, over 90% of the Arkansas rice production area utilizes a delayed-flood system, where the permanent flood is not established until the four- to five-leaf growth stage, which generally occurs approximately three to four weeks after emergence [39]. Fields are drained two to three weeks prior to harvest and most fields remain unflooded until the subsequent rice crop is produced, while nearly 20% of Arkansas rice area is winter-flooded [34,39]. Over 75% of Arkansas rice is irrigated by groundwater with 10 and 13% of the rice area utilizing water stored in reservoirs and from streams/rivers, respectively [39].

The two methods of nitrogen (N) fertilization in Arkansas are (i) the standard two-way split system, where 65 to 75% of the total N is applied pre-flood with the remainder applied at midseason in one or two applications between beginning internode elongation and half-inch internode elongation [i.e., reproductive stage 0 (R0) to 1 (R1)], and (ii) the single optimum preflood system, where a single N application is made immediately prior to flooding. Nitrogen fertilizer rate recommendations have previously been based only on cultivar, soil texture, and previous crop. Implementation of the new N-Soil Test for Rice (N-STaR) enables recommen‐ dations to be adapted to the soil's ability to supply N to the rice crop on a field-by-field basis, reducing the likelihood of over- and under-fertilization of N [43]. Ammonium-N sources, such as urea and ammonium sulfate, are used in order to prevent N loss through denitrification that occurs with nitrate-containing fertilizers. Phosphorus and potassium are incorporated prior to planting as recommended by routine soil tests [43]. Organic amendments are uncommon, although poultry litter is utilized to a small degree, especially in precision-leveled fields.

#### **5. Flooded soils**

**4.4. Arkansas rice production practices**

186 Greenhouse Gases

early- to mid-September [31].

poor seed-to-soil contact [41].

streams/rivers, respectively [39].

Arkansas is the leading rice-producing state, accounting for 40 to 50% of total annual produc‐ tion in the US [11]. Rice production in Arkansas began in 1902 when 0.4 ha were planted in Lonoke County. Production increased over time until 1955 when government quotas limited production to 202,350 ha. The limitation was lifted in 1974 and production increased again, peaking in 1981 at 623,240 ha, again in 1999 with 667,755 harvested hectares, and finally in 2010 with 724,413 ha [31]. In 2012, 518,016 ha rice were harvested in Arkansas [11]. Rice production in Arkansas is highly mechanized with a heavy dependence upon synthetic fertilizers, chemical pest control, and machinery. Planting of rice in Arkansas generally begins the last week of March and extends into early June with floods typically being established by the end of May or early June. Harvesting operations usually begin in mid-August and peak in

Arkansas rice is produced on a wide variety of soils ranging from sandy to clay soils with the differing textural classes generally requiring different management, especially with regards to tillage practices and nutrient management [39, 41]. Production on sands and sandy loams is minor and has been decreasing from 3.1 and 5.2% of Arkansas area, respectively, in 2007 to 0.7 and 3.7%, respectively, in 2012. Arkansas production on clay and clay-loam soils, however, has increased from under 40 to 48% between 2007 and 2009 but declined to 43% in 2012. Production on silt-loam soils has remained fairly steady at 52% in 2007 and 53% in 2012 [39,42].

Dry-seeding techniques have always dominated in Arkansas. Water-seeding has varied between 2 and 8% of the production area between 2007 and 2012, with an estimated 5% of the 2012 Arkansas rice area being water-seeded [39,42]. Approximately 80% of 2012 Arkansas rice area was drill-seeded, compared to approximately 20% being broadcast-seeded [39]. Conven‐ tional tillage accounted for over half of Arkansas planted-rice area, while stale-seedbed (i.e., tillage and floating, or leveling the field, in the fall or winter) and no-tillage accounted for 35 and 10% of the planted-rice area, respectively, in 2012 [39]. Stale-seedbed and no-tillage are oftentimes utilized on clay soils where conventional tillage produces a cloddy seedbed with

While pinpoint, water-seeding techniques do occur in Arkansas, over 90% of the Arkansas rice production area utilizes a delayed-flood system, where the permanent flood is not established until the four- to five-leaf growth stage, which generally occurs approximately three to four weeks after emergence [39]. Fields are drained two to three weeks prior to harvest and most fields remain unflooded until the subsequent rice crop is produced, while nearly 20% of Arkansas rice area is winter-flooded [34,39]. Over 75% of Arkansas rice is irrigated by groundwater with 10 and 13% of the rice area utilizing water stored in reservoirs and from

The two methods of nitrogen (N) fertilization in Arkansas are (i) the standard two-way split system, where 65 to 75% of the total N is applied pre-flood with the remainder applied at midseason in one or two applications between beginning internode elongation and half-inch internode elongation [i.e., reproductive stage 0 (R0) to 1 (R1)], and (ii) the single optimum preflood system, where a single N application is made immediately prior to flooding. Nitrogen The saturated soils that occur during wetland, or lowland, rice cultivation give rise to a set of physical, chemical, and biological properties that are quite different from upland soils. Rice is the only major row crop produced under flooded-soil conditions and the absence of air-filled pores along with reduced soil–atmosphere interactions result in an almost entirely different set of processes than those occurring in upland cropping systems.

#### **5.1. Physical characteristics of flooded soils**

The major physical difference between saturated and unsaturated soils involves the availabil‐ ity and rates of movement for gases and solutes. Under aerated conditions, the soil atmosphere contains essentially the same gases as the atmosphere although the proportions of oxygen (O2) and CO2 differ from the atmosphere due to soil respiration [44]. Carbon dioxide diffuses into the atmosphere from the soil due to production during respiration and O2 diffuses into the soil as it is consumed during respiration. The saturation and ponding of flooded soils greatly reduce gas transport between the soil and atmosphere compared to aerated soils and plant-mediated transport of gases by diffusion is often the main exchange mechanism between the soil and atmosphere in saturated or flooded systems [45]. As a flooded soil dries, gases trapped in the soil may escape due to increases in diffusion and convective flow rates that occur as water escapes soil pores.

While solute movement by diffusion may be greater in saturated soils due to an increase in water-filled pore space, diffusion of gases through water is roughly three to four orders of magnitude slower than diffusion of gases through air [46,47]. Both diffusive and convective flow of gases and solutes are related to pore connectivity and tortuosity, so it is expected that movement of gases and solutes are slower in fine-textured soils, such as clays and clay loams, than in coarser-textured soils, such as silt loams and sands, which generally have larger, more connected pores [47]. Convective flow of gases in saturated soils can occur as dissolved gases move with moving soil water, which is dependent largely upon soil texture and structure, and as ebullition, which is where gases escape as bubbles through ponded water [47]. Generally, diffusion dominates gas transport in fine-textured soils, such as clay loams and clays, and diffusion rates typically decrease as particle size decreases, which is due to differences in size, orientation, and shape of soil pore spaces [45,48]. Soil texture also affects the amount of time it takes for a soil to become saturated with infiltration rates in clayey soils estimated to be 1 to 5 mm hr–1 compared to 10 to 20 mm hr–1 in soils such as silt loams [47]. The amount of time a soil takes to become saturated has an effect on chemical and biological processes that develop as the system becomes anaerobic.

#### **5.2. Soil redox potential**

Isolation of flooded soils from the atmosphere and depletion of soil O2 induces biological and chemical reactions that create anaerobic and reducing conditions rather than the aerobic and oxidized conditions that generally occur in upland soils. Organic matter decomposition slows under anaerobic conditions, but as organic matter is oxidized, transformations such as denitrification and manganese (Mn) and iron (Fe) reduction occur as well as production of gases such as hydrogen sulfide (H2S) and CH4. Soil reduction/oxidation (redox) reactions are coupled half-reactions where the oxidation of organic matter, which provides electrons, is coupled with the reduction of elements or compounds that act as electron acceptors [49]. Oxygen is the major electron acceptor under aerobic conditions, but as O2 is depleted, the sequence of electron acceptors shifts to NO3 – , MnO2, Fe(OH)3, SO4 2–, and CO2, which are theoretically reduced in that order based on thermodynamic favorability [44,50]. The reduced forms of the previously mentioned terminal electron acceptors are H2O, N2, Mn2+, Fe2+, H2S, and CH4, respectively. Soil redox reactions in a controlled laboratory environment may follow the theoretical sequence, but environmental conditions in the field result in spatial variability of oxidizable organic compounds, electron acceptors, and microorganisms that cause sub‐ stantial overlap of the terminal electron acceptor sequence [44,49].

Soil redox potential (Eh) is a measure of the electrical potential status of a system that results from the tendency of substances in the system to donate or acquire electrons [51]. Soil redox potential is measured in millivolts (mV) using a platinum electrode along with a mercury chloride (HgCl) or silver chloride (AgCl) reference electrode, both connected to a voltmeter [49]. Combination platinum electrodes are also available that can continuously monitor soil Eh when connected to a logger box. When using AgCl electrodes, a correction factor of approximately +200 mV is added to field-measured voltages in order to adjust measurements to the standard hydrogen electrode [52]. In well-aerated soils, soil Eh may be as great as +700 mV, but Eh values near –300 mV may be observed in saturated organic-matter-rich soils [51]. As a system shifts from aerobic to anaerobic and soil redox potential declines, atmospheric O2 is reduced first at +380 to +320 mV, followed by NO3 – (+280 to +220 mV), MnO2 (+220 to +180 mV), Fe(OH)3 (+110 to +80 mV), SO4 2– (–140 to –170 mV), and CO2 (–200 to –280 mV), based on measurements by Patrick and Jugsujinda [53].

#### **6. Methane emissions from rice**

Methane emissions from any ecosystem, particularly a rice agroecosystem (Figure 1), are governed by the magnitude and balance of microbial CH4 production (methanogenesis) and oxidation (methanotrophy), which occur by separate microbial communities. The two groups of microorganisms are adapted to different environmental conditions, and, as a result, are affected differently based on the structure and conditions of an ecosystem, which results in variability of CH4 production and oxidation potentials across time and space [54]. With low CH4productionratesorlongdiffusionpathways,it seems thatthemajorityoftheCH4produced is oxidized. Conversely, in cases where CH4 production rates are high or diffusion paths are short, less CH4 is oxidized and a greater portion reaches the atmosphere [54] (Figure 1).

it takes for a soil to become saturated with infiltration rates in clayey soils estimated to be 1 to 5 mm hr–1 compared to 10 to 20 mm hr–1 in soils such as silt loams [47]. The amount of time a soil takes to become saturated has an effect on chemical and biological processes that develop

Isolation of flooded soils from the atmosphere and depletion of soil O2 induces biological and chemical reactions that create anaerobic and reducing conditions rather than the aerobic and oxidized conditions that generally occur in upland soils. Organic matter decomposition slows under anaerobic conditions, but as organic matter is oxidized, transformations such as denitrification and manganese (Mn) and iron (Fe) reduction occur as well as production of gases such as hydrogen sulfide (H2S) and CH4. Soil reduction/oxidation (redox) reactions are coupled half-reactions where the oxidation of organic matter, which provides electrons, is coupled with the reduction of elements or compounds that act as electron acceptors [49]. Oxygen is the major electron acceptor under aerobic conditions, but as O2 is depleted, the

–

theoretically reduced in that order based on thermodynamic favorability [44,50]. The reduced forms of the previously mentioned terminal electron acceptors are H2O, N2, Mn2+, Fe2+, H2S, and CH4, respectively. Soil redox reactions in a controlled laboratory environment may follow the theoretical sequence, but environmental conditions in the field result in spatial variability of oxidizable organic compounds, electron acceptors, and microorganisms that cause sub‐

Soil redox potential (Eh) is a measure of the electrical potential status of a system that results from the tendency of substances in the system to donate or acquire electrons [51]. Soil redox potential is measured in millivolts (mV) using a platinum electrode along with a mercury chloride (HgCl) or silver chloride (AgCl) reference electrode, both connected to a voltmeter [49]. Combination platinum electrodes are also available that can continuously monitor soil Eh when connected to a logger box. When using AgCl electrodes, a correction factor of approximately +200 mV is added to field-measured voltages in order to adjust measurements to the standard hydrogen electrode [52]. In well-aerated soils, soil Eh may be as great as +700 mV, but Eh values near –300 mV may be observed in saturated organic-matter-rich soils [51]. As a system shifts from aerobic to anaerobic and soil redox potential declines, atmospheric

Methane emissions from any ecosystem, particularly a rice agroecosystem (Figure 1), are governed by the magnitude and balance of microbial CH4 production (methanogenesis) and oxidation (methanotrophy), which occur by separate microbial communities. The two groups

–

, MnO2, Fe(OH)3, SO4

2–, and CO2, which are

(+280 to +220 mV), MnO2 (+220 to +180

2– (–140 to –170 mV), and CO2 (–200 to –280 mV), based on

as the system becomes anaerobic.

sequence of electron acceptors shifts to NO3

stantial overlap of the terminal electron acceptor sequence [44,49].

O2 is reduced first at +380 to +320 mV, followed by NO3

measurements by Patrick and Jugsujinda [53].

mV), Fe(OH)3 (+110 to +80 mV), SO4

**6. Methane emissions from rice**

**5.2. Soil redox potential**

188 Greenhouse Gases

2 Figure 1. Chamber-based measurements of methane emissions from small plots at the Rice 3 Research and Extension Center near Stuttgart, AR (top), and at the Northeast Research 4 and Extension Center at Keiser, AR (bottom). Photographs taken by K. Brye. **Figure 1.** Chamber-based measurements of methane emissions from small plots at the Rice Research and Extension Center near Stuttgart, AR (top), and at the Northeast Research and Extension Center at Keiser, AR (bottom). Photo‐ graphs taken by K. Brye.

1

43

#### **6.1. Methane production and oxidation**

Methane production occurs toward the end of a complex anaerobic decomposition process in which soil organic matter (SOM) is degraded to acetate, hydrogen gas (H2), and CO2 by a community of various fermenting microorganisms, which are mostly bacteria. Methanogenic *Archaea* are then able to split acetate into CH4 and CO2 (i.e., acetoclastic methanogenesis) or utilize H2 and CO2 to produce CH4 (i.e., hydrogenotrophic methanogenesis) [55,56]. Metha‐ nogens encompass a large group of strictly anaerobic, obligate *Archaea*, which is currently composed of three classes, six orders, 12 families, and 35 genera [56]. Rice Cluster I is a specific group of methanogens identified by Grosskopf et al. [57] that contains enzymes in order to detoxify highly reactive O2 species, allowing the methanogens to survive in aerated soils or oxygenated rhizospheres, and occurs preferentially in environments that undergo transient aerobic conditions, such as in rice fields [55,58]. Rice Cluster I has been detected in almost all rice field soils tested [59,60] and occurs in great abundance in rice soils and on rice roots, representing up to 50% of total methanogens in rice fields [61]. Rice Cluster I has been identified as occupying a niche on rice roots by producing CH4 from photosynthates released as root exudates [55,62]. Recent research has confirmed that methanogens are ubiquitous in aerobic soils and have the ability to produce CH4 as soon as anoxic conditions form and substrate is available [56]. Conrad [63] reported that methanogens isolated from the soil of rice fields were not killed but only inhibited by high redox potentials or O2 exposure, allowing them to survive drainage and maintain their population size throughout the year in a state of low activity.

Most methanogens are mesophiles and neutrophiles, with optimal growth occurring between 30 and 40o C and between a pH of 6 and 8 [54]. Methanogens are highly sensitive to variations in temperature and pH and CH4 production is greatly reduced when soil temperatures are low or in acidic or alkaline soils [56]. Within the optimal temperature range, which is generally the case during the rice growing season, temperature has a positive effect on methanogenesis, causing an increase in CH4 production as temperature increases [54,56].

Methane oxidation is achieved by a group of aerobic *Proteobacteria* known as methanotrophs, which only utilize CH4 or methanol as a source of C and energy and are currently classified into two phyla, three orders, four families, 21 genera, and 56 species [56]. One group, known as low-affinity methanotrophs, is capable of oxidizing high CH4 concentrations (>100 ppm) and exists at oxic–anoxic interfaces, where the methanotrophs consume CH4 produced in anoxic environments [56]. Another group, known as high-affinity methanotrophs, exists in upland soils and possesses the ability to oxidize CH4 at low atmospheric levels (<2 ppm) [64]. Unlike methanogenesis, methanotrophy is not impacted greatly by temperature, although CH4 oxidation is decreased below 10o C and above 40o C, or pH, as similar CH4 oxidation has been observed in soils with pH values ranging from 3.5 to 8 [56]. Due to the differing effect of temperature on methanogenesis and methanotrophy, CH4 production increases as soil temperatures increase, while CH4 oxidation changes little, resulting in a general increase in CH4 emissions as soil temperature increases throughout the rice growing season. This effect has been confirmed in a laboratory incubation of anaerobic soils at various temperatures between 5 and 25o C [65].

#### **6.2. Substrate for methane production**

**6.1. Methane production and oxidation**

30 and 40o

190 Greenhouse Gases

CH4 oxidation is decreased below 10o

C [65].

between 5 and 25o

Methane production occurs toward the end of a complex anaerobic decomposition process in which soil organic matter (SOM) is degraded to acetate, hydrogen gas (H2), and CO2 by a community of various fermenting microorganisms, which are mostly bacteria. Methanogenic *Archaea* are then able to split acetate into CH4 and CO2 (i.e., acetoclastic methanogenesis) or utilize H2 and CO2 to produce CH4 (i.e., hydrogenotrophic methanogenesis) [55,56]. Metha‐ nogens encompass a large group of strictly anaerobic, obligate *Archaea*, which is currently composed of three classes, six orders, 12 families, and 35 genera [56]. Rice Cluster I is a specific group of methanogens identified by Grosskopf et al. [57] that contains enzymes in order to detoxify highly reactive O2 species, allowing the methanogens to survive in aerated soils or oxygenated rhizospheres, and occurs preferentially in environments that undergo transient aerobic conditions, such as in rice fields [55,58]. Rice Cluster I has been detected in almost all rice field soils tested [59,60] and occurs in great abundance in rice soils and on rice roots, representing up to 50% of total methanogens in rice fields [61]. Rice Cluster I has been identified as occupying a niche on rice roots by producing CH4 from photosynthates released as root exudates [55,62]. Recent research has confirmed that methanogens are ubiquitous in aerobic soils and have the ability to produce CH4 as soon as anoxic conditions form and substrate is available [56]. Conrad [63] reported that methanogens isolated from the soil of rice fields were not killed but only inhibited by high redox potentials or O2 exposure, allowing them to survive drainage and maintain their population size throughout the year in a state of low activity.

Most methanogens are mesophiles and neutrophiles, with optimal growth occurring between

in temperature and pH and CH4 production is greatly reduced when soil temperatures are low or in acidic or alkaline soils [56]. Within the optimal temperature range, which is generally the case during the rice growing season, temperature has a positive effect on methanogenesis,

Methane oxidation is achieved by a group of aerobic *Proteobacteria* known as methanotrophs, which only utilize CH4 or methanol as a source of C and energy and are currently classified into two phyla, three orders, four families, 21 genera, and 56 species [56]. One group, known as low-affinity methanotrophs, is capable of oxidizing high CH4 concentrations (>100 ppm) and exists at oxic–anoxic interfaces, where the methanotrophs consume CH4 produced in anoxic environments [56]. Another group, known as high-affinity methanotrophs, exists in upland soils and possesses the ability to oxidize CH4 at low atmospheric levels (<2 ppm) [64]. Unlike methanogenesis, methanotrophy is not impacted greatly by temperature, although

C and above 40o

been observed in soils with pH values ranging from 3.5 to 8 [56]. Due to the differing effect of temperature on methanogenesis and methanotrophy, CH4 production increases as soil temperatures increase, while CH4 oxidation changes little, resulting in a general increase in CH4 emissions as soil temperature increases throughout the rice growing season. This effect has been confirmed in a laboratory incubation of anaerobic soils at various temperatures

C, or pH, as similar CH4 oxidation has

causing an increase in CH4 production as temperature increases [54,56].

C and between a pH of 6 and 8 [54]. Methanogens are highly sensitive to variations

Available SOM stimulates CH4 production due to enhanced fermentative production of acetate and H2/CO2 and, in principle, CH4 production could be expected to be proportional to organic C inputs, but the reduction of nitrate (NO3 - ), iron (Fe), manganese (Mn), and sulfate (SO4 2-) all precede methanogenesis and reduce the amount of available C for CH4 production [54]. Methane production may be stimulated by root exudates [66–68] or the application of animal manures [69], green manures [70–73], or rice straw [67,70,73–75], while the application of composted organic C sources does not greatly increase CH4 production [73,75,76]. This indicates that the amount of available organic C (OC) is more important in determining CH4 production than total OC (TOC), as composted residue contains lower amounts of degradable C, on a mass basis, compared to fresh residues [77]. Yagi and Minami [73] and Wang et al. [78] confirmed a positive correlation between CH4 production and readily mineralizable C, while studies have indicated no clear relationship between soil TOC and CH4 production [68,79– 81]. Research conducted by Denier van der Gon and Neue [76] determined that increasing fresh OM inputs would result in increases in CH4 production up to a point where another factor becomes limiting; however, fresh green manure inputs up to 20 Mg ha–1 still indicated OC limitations. In most rice production situations, organic residue inputs are below 20 Mg ha–1 and will generally exhibit an increase in CH4 emissions as organic inputs increase.

Using 13C-labeled rice straw incorporated at 6 Mg ha–1, Watanabe et al. [82] determined that 42% of season-long CH4 emissions originated from rice straw C, 37 to 40% from the rice plant, and 18 to 21% from SOM. The contribution of SOM to CH4 production was fairly consistent over the growing season, while the contribution from rice straw decreased from nearly 90% at 14 days after transplanting to only 11 to 16% during heading and grain fill. In contrast, the contribution of living rice plants to CH4 production increased over time and amounted to 65 to 70% during heading and grain fill [82]. Chidthaisong and Watanabe [83] also observed that the contribution of rice straw to CH4 production was greatest at 20 to 40 days after flooding, while plant-derived C became increasingly more influential as the season progressed. The link between root exudates and CH4 production has been observed directly by Aulakh et al. [84], who showed a positive correlation between TOC in root exudates and CH4 production. Several others have observed an inverse relationship between grain yield and CH4 production [19,85], indicating that lower grain yields are accompanied by greater CH4 production as a result of greater root exudation, which was confirmed by Aulakh et al. [66]. Using 13C-labeled CO2, it was observed that photosynthates were a major source of CH4 and accounted for 4 to 52% of CH4 under field conditions [86,87].

#### **6.3. Duration and timing of methane production**

Methane production occurs for some period of time following a period of prolonged saturated conditions and continues until the C substrate becomes limiting or environmental conditions limit methanogenesis (i.e., the soil becomes too cold, hot, or aerated). In flooded soils, the rate of reduction processes is determined by the composition and texture of a soil as well as the content of inorganic electron acceptors [i.e., NO3 – , MnO2, Fe(OH)3, SO4 2–] and available C, so the amount of time between flooding a soil and the onset of methanogenesis can vary from several days to several weeks [88]. From the onset of methanogenesis, CH4 emissions from rice systems generally increase over time as the soil becomes more reduced and usually shows one or more of three general peak flux trends. Early season peak fluxes are generally attributed to decomposition of freshly incorporated residues and generally occur within 20 to 40 days after flooding [83,89] and late-season peaks are thought to result from decomposition following senescence of rice roots [90,91]. The other time period of peak fluxes generally occurs near the time of 50% heading (i.e., approximately the time of anthesis) and has been linked to the sinksource relationship of photosynthates in the plant when CH4 fluxes have been observed to increase during vegetative growth as root exudates increase and decrease following heading as fixed-C is translocated to developing grain. This plant-related peak has been observed in several studies [15–17,80,92–94,95] and similar seasonal trends have been observed in root growth [96–98], root exudation rates [66], and anaerobic root respiration rates [99].

#### **6.4. Transport mechanisms**

The three mechanisms by which CH4 is transported from a ponded soil to the atmosphere are diffusion through the floodwater, ebullition, and plant-mediated diffusion. Diffusion of CH4 through overlying floodwater is minor as diffusion of gases is approximately 10,000 times slower through water than through air [46]. Ebullition, bubbles forming and forcing their way to the surface, may be a significant transport mechanism early in the season, especially with high OM inputs, soil disturbances, and in coarse-textured soils, but generally plays only a small role in CH4 transport, which diminishes as plants mature and plant-mediated transport (PMT) increases [76,100]. The majority of CH4 emissions from a rice system occur through the rice plants via aerenchyma cells, where studies have indicated that about 90% of season-long emissions are released through the rice plants, compared to 8 to 9% released by ebullition and 1 to 2% by diffusion through the floodwater [100–104].

Based on experiments using artificial atmospheres of various gas compositions, Denier van der Gon and van Breemen [105] determined that PMT is driven by molecular diffusion and not affected by transpiration or stomatal opening. Others have observed a decreasing CH4 concentration gradient from the soil to the rice root aerenchyma, shoot aerenchyma, and atmosphere, indicative of a diffusive transport pathway from the soil to the atmosphere through the plant [104,106]. Other studies have also confirmed that CH4 transport is not related to transpiration and is unaffected by cutting plants just above the water surface [103,104,107]. However, Hosono and Nouchi [108] determined that PMT was reduced linearly as roots were cut and increased with root growth up to heading, indicating that the surface area of roots in contact with soil solution is important in determining PMT. Several studies have determined that the most restrictive zone of CH4 transport through the rice plant is the root–shoot transition zone where dense intercalary meristem cells restrict movement from the root aerenchyma to the shoot aerenchyma [101,105,106,109,110].

It has been postulated that CH4 in the gaseous form or dissolved in water enters into root aerenchyma, which forms by degeneration of cortical cells between the exodermis and the vascular bundle, where the dissolved CH4 is gasified and moves by diffusion from the root aerenchyma through the restrictive transition zone into the aerenchyma of the culm and then to the atmosphere [104,107,109]. It has been determined that CH4 is released from the rice plant mainly through the lower leaf sheaths. Examining the cultivar 'Koshihikari' with a scanning electron microscope, Nouchi et al. [104] and Nouchi and Mariko [107] observed CH4 release from 4-µm diameter, hook-shaped micropores arranged regularly approximately 80 µm apart on the abaxial epidermis of leaf sheaths as well as from the connections of leaf sheaths to the culm at nodes. Butterbach-Bahl et al. [106] also determined that CH4 is primarily released through the lower leaf sheaths, however, micropores were not observed in the cultivars 'Roma' or 'Lido'. More research is required to determine differences in CH4 release from various cultivars. It has been determined that rice cultivars have differences in CH4 transport capacity, likely in relation to differences in aerenchyma morphology and the root–shoot transition zone [101] and that CH4 transport capacity increases as soil temperature increases [108]. Research indicates that PMT is the dominant mechanism of CH4 release from rice soils and that the rate of transport can be influenced by cultivar or environmental conditions.

#### **7. Factors affecting methane emissions from rice**

several days to several weeks [88]. From the onset of methanogenesis, CH4 emissions from rice systems generally increase over time as the soil becomes more reduced and usually shows one or more of three general peak flux trends. Early season peak fluxes are generally attributed to decomposition of freshly incorporated residues and generally occur within 20 to 40 days after flooding [83,89] and late-season peaks are thought to result from decomposition following senescence of rice roots [90,91]. The other time period of peak fluxes generally occurs near the time of 50% heading (i.e., approximately the time of anthesis) and has been linked to the sinksource relationship of photosynthates in the plant when CH4 fluxes have been observed to increase during vegetative growth as root exudates increase and decrease following heading as fixed-C is translocated to developing grain. This plant-related peak has been observed in several studies [15–17,80,92–94,95] and similar seasonal trends have been observed in root

growth [96–98], root exudation rates [66], and anaerobic root respiration rates [99].

The three mechanisms by which CH4 is transported from a ponded soil to the atmosphere are diffusion through the floodwater, ebullition, and plant-mediated diffusion. Diffusion of CH4 through overlying floodwater is minor as diffusion of gases is approximately 10,000 times slower through water than through air [46]. Ebullition, bubbles forming and forcing their way to the surface, may be a significant transport mechanism early in the season, especially with high OM inputs, soil disturbances, and in coarse-textured soils, but generally plays only a small role in CH4 transport, which diminishes as plants mature and plant-mediated transport (PMT) increases [76,100]. The majority of CH4 emissions from a rice system occur through the rice plants via aerenchyma cells, where studies have indicated that about 90% of season-long emissions are released through the rice plants, compared to 8 to 9% released by ebullition and

Based on experiments using artificial atmospheres of various gas compositions, Denier van der Gon and van Breemen [105] determined that PMT is driven by molecular diffusion and not affected by transpiration or stomatal opening. Others have observed a decreasing CH4 concentration gradient from the soil to the rice root aerenchyma, shoot aerenchyma, and atmosphere, indicative of a diffusive transport pathway from the soil to the atmosphere through the plant [104,106]. Other studies have also confirmed that CH4 transport is not related to transpiration and is unaffected by cutting plants just above the water surface [103,104,107]. However, Hosono and Nouchi [108] determined that PMT was reduced linearly as roots were cut and increased with root growth up to heading, indicating that the surface area of roots in contact with soil solution is important in determining PMT. Several studies have determined that the most restrictive zone of CH4 transport through the rice plant is the root–shoot transition zone where dense intercalary meristem cells restrict movement from the root aerenchyma to

It has been postulated that CH4 in the gaseous form or dissolved in water enters into root aerenchyma, which forms by degeneration of cortical cells between the exodermis and the vascular bundle, where the dissolved CH4 is gasified and moves by diffusion from the root aerenchyma through the restrictive transition zone into the aerenchyma of the culm and then

**6.4. Transport mechanisms**

192 Greenhouse Gases

1 to 2% by diffusion through the floodwater [100–104].

the shoot aerenchyma [101,105,106,109,110].

Through numerous research efforts since the 1980s, several factors have been determined to affect CH4 emissions from rice cultivation. Due to the complex balance of methanogenesis and methanotrophy that determines how much CH4 escapes the rice system to the atmosphere along with the large variety of cultural and environmental conditions around the globe, there is large variability in the impact of different factors across time and space. There are a few soil, environmental, and plant factors, however, that seem to have somewhat consistent impacts on CH4 emissions from rice.

#### **7.1. Soil factors affecting methane emissions from rice**

Various studies have observed inconsistent results of N fertilizer application on CH4 emissions including an increase in emissions with added N [85,90,111,112], a decrease in emissions with added N [113,114], or no impact of added N on CH4 emissions [15,75,115]. Banger et al. [116] conducted a meta-analysis and determined that CH4 emissions were significantly greater from N-fertilized rice in 98 out of 155 data pairs, indicating that the increase in plant growth and C fixation resulting from N-fertilization generally increases CH4 emissions. Wang et al. [78] postulated that the effect of urea on CH4 emissions may be impacted by pH, where it was observed that urea may cause a decrease in emissions in alkaline soils as urea hydrolysis increases soil pH, limiting the neutrophilic methanogens. In acidic soils, however, the increase in pH from urea hydrolysis shifts the soil pH toward neutral and enhances methanogenesis. Research has consistently indicated that ammonium sulfate reduces CH4 emissions relative to urea application [70,113,116], likely due to the impact of soil acidification and sulfate reduction decreasing the available C substrate for methanogenesis. Similarly, other studies have determined that oxidized Fe [80,117–120] or NO3 – [120] amendments have the ability to reduce CH4 emissions. In addition, Lu et al. [121] observed a 19 to 33% reduction in CH4 emissions with the application of P due to enhanced root growth and root exudation that was measured in the P-deficient treatment.

Multiple studies have indicated no significant correlations between CH4 emissions and any static soil properties [68,81] or between CH4 emissions and total soil C [79,80], while readily mineralizable C has been shown to be positively correlated with CH4 emissions [75,78]. Particle-size distribution is one soil property that has been regularly related to CH4 emissions as emissions have been positively correlated with soil sand content [78,80,118,119,122] and inversely correlated with soil clay content [71,78,118,119,122,123]. Studies have observed an increase in CH4 entrapment resulting from increasing clay contents [71,78], and Sass and Fisher [91] attributed the reduction in CH4 emissions from clay soils to the entrapment and slow movement of CH4 that allows more CH4 to be oxidized in aerated zones surrounding roots and at the soil surface. In a laboratory incubation study, Wang et al. [78] observed varying degrees of CH4 entrapment, even among soils with similar sand and clay contents, where the greatest entrapment (98%) was measured from a Sharkey clay (very-fine, smectitic, thermic Chromic Epiaquerts) soil compared to 81 and 68% entrapment from a Beaumont clay (fine, smectitic, hyperthermic Chromic Dystraquerts) and a Sacramento clay (very-fine, smectitic, thermic Cumulic Vertic Endoaquolls), respectively. This research indicates that clayey soils have the capability of restricting movement of CH4 to the atmosphere and that other factors, such as clay minerology and soil chemical properties, may impact emissions more than simply the total amount of clay.

#### **7.2. Environmental factors affecting methane emissions from rice**

Two major environmental factors that impact CH4 emissions from rice are temperature and soil saturation status. Numerous studies have observed increases in CH4 fluxes in relation to increasing soil temperatures [100,108,124]. A study conducted in Japan observed a 1.6-fold increase in emissions from one year to another under the same management and location resulting from an increase in average air temperature from 24.6 to 26.9o C [119]. Methanotrophic activity changes only slightly between 10 and 40o C, while temperature has a strong influence on methanogenesis [56], which leads to a decrease in the proportion of CH4 oxidized and an increase in emissions as soil temperature increases. Van Winden et al. [65], for example, reported 98% CH4 oxidation at 5o C compared to 50% oxidation at 25o C.

Soil saturation status has a profound influence on CH4 emissions through the impact of saturation on soil redox processes, such as methanogenesis. Methane emissions have been observed from soils at an Eh as great as –100 mV [125], while emissions increase as Eh decreases. The amount of time required after saturation to reach low redox potentials condu‐ cive to methanogenesis varies based on soil textural and chemical properties [119], but generally occurs within several days or weeks after flooding. Studies have indicated that a single mid-season drainage can reduce CH4 emissions by as much as 65% [68,70,75,95,113,126], however, the potential for greenhouse gas mitigation is reduced or negated due to an increase in N2O emissions resulting from the drainage [70,113,126,127]. Further research is needed in order to more adequately understand the balance between CH4 and N2O emissions under various water management regimes as well as the impact that N management has on emissions when fields are drained.

#### **7.3. Plant factors affecting methane emissions from rice**

Multiple studies have indicated no significant correlations between CH4 emissions and any static soil properties [68,81] or between CH4 emissions and total soil C [79,80], while readily mineralizable C has been shown to be positively correlated with CH4 emissions [75,78]. Particle-size distribution is one soil property that has been regularly related to CH4 emissions as emissions have been positively correlated with soil sand content [78,80,118,119,122] and inversely correlated with soil clay content [71,78,118,119,122,123]. Studies have observed an increase in CH4 entrapment resulting from increasing clay contents [71,78], and Sass and Fisher [91] attributed the reduction in CH4 emissions from clay soils to the entrapment and slow movement of CH4 that allows more CH4 to be oxidized in aerated zones surrounding roots and at the soil surface. In a laboratory incubation study, Wang et al. [78] observed varying degrees of CH4 entrapment, even among soils with similar sand and clay contents, where the greatest entrapment (98%) was measured from a Sharkey clay (very-fine, smectitic, thermic Chromic Epiaquerts) soil compared to 81 and 68% entrapment from a Beaumont clay (fine, smectitic, hyperthermic Chromic Dystraquerts) and a Sacramento clay (very-fine, smectitic, thermic Cumulic Vertic Endoaquolls), respectively. This research indicates that clayey soils have the capability of restricting movement of CH4 to the atmosphere and that other factors, such as clay minerology and soil chemical properties, may impact emissions more than simply

Two major environmental factors that impact CH4 emissions from rice are temperature and soil saturation status. Numerous studies have observed increases in CH4 fluxes in relation to increasing soil temperatures [100,108,124]. A study conducted in Japan observed a 1.6-fold increase in emissions from one year to another under the same management and location

on methanogenesis [56], which leads to a decrease in the proportion of CH4 oxidized and an increase in emissions as soil temperature increases. Van Winden et al. [65], for example,

Soil saturation status has a profound influence on CH4 emissions through the impact of saturation on soil redox processes, such as methanogenesis. Methane emissions have been observed from soils at an Eh as great as –100 mV [125], while emissions increase as Eh decreases. The amount of time required after saturation to reach low redox potentials condu‐ cive to methanogenesis varies based on soil textural and chemical properties [119], but generally occurs within several days or weeks after flooding. Studies have indicated that a single mid-season drainage can reduce CH4 emissions by as much as 65% [68,70,75,95,113,126], however, the potential for greenhouse gas mitigation is reduced or negated due to an increase in N2O emissions resulting from the drainage [70,113,126,127]. Further research is needed in order to more adequately understand the balance between CH4 and N2O emissions under various water management regimes as well as the impact that N management has on emissions

C compared to 50% oxidation at 25o

C [119]. Methanotrophic

C, while temperature has a strong influence

C.

the total amount of clay.

194 Greenhouse Gases

**7.2. Environmental factors affecting methane emissions from rice**

resulting from an increase in average air temperature from 24.6 to 26.9o

activity changes only slightly between 10 and 40o

reported 98% CH4 oxidation at 5o

when fields are drained.

Due to the strong impact of rice plants on CH4 transport and CH4 production from root exudates and residue, there are several plant factors that significantly impact emissions from rice cultivation. A strong relationship between plant growth and CH4 emissions has been observed in many studies [16,17,80,92–95], particularly in temperate regions, where much of the previous crop's residue decomposes during the winter. Studies have indicated that CH4 emissions are up to 20 times greater from soil planted with rice than from unvegetated soil [67,107,123], indicating the large influence of rice plants on emissions.

One of the major plant factors impacting CH4 emissions from rice is whether or not a ratoon crop is grown. This impact is reflected in the USEPA's emissions factors, which are 178 kg CH4-C ha–1 for non-California primary rice crops and an additional 585 kg CH4-C ha–1 when a ratoon crop is produced [2], based on ratoon crops studied in Louisiana [21,22]. The large increase in emissions from ratoon crops is likely a result of large quantities of residue inputs from the harvest of the primary crop in addition to well-developed root systems that further increase the available C for methanogenesis. Lindau et al. [22] observed a significant positive correlation between rice straw additions from a primary crop and resulting emissions from the following ratoon crop.

Another plant factor that has a substantial impact on CH4 emissions is biomass accumulation. Huang et al. [128] determined that CH4 fluxes measured during the growing season were positively correlated to aboveground and belowground dry matter on the dates of flux measurements. Additional studies have observed positive correlations between season-long CH4 emissions and aboveground [16,72,102,128] and belowground dry matter [129]. These studies have indicated a strong relationship between plant growth and CH4 emissions, which may result from an increase in available substrate as root exudates have been correlated to biomass [66].

Cultivar selection has also been shown to be an important plant factor influencing CH4 emissions from rice. While the mechanisms for cultivar differences in CH4 emissions have not been extensively studied, it appears that differences likely arise from variability in CH4 transport capacity, biomass or dry matter production, root exudation, and microbial com‐ munity dynamics among cultivars. Butterbach-Bahl et al. [101], for example, attributed a 24 to 31% difference in emissions between two pure-line cultivars to differences in CH4 transport capacities, as no differences were observed between CH4 production or oxidation. Aulakh et al. [84] observed a positive correlation between TOC from root exudates and CH4 production potential, indicating the potential for cultivar differences in emissions based on variable root exudation rates. Previous studies have reported reduced emissions from semi-dwarf relative to standard-stature cultivars [22,91,130]. The difference in CH4 emissions between semi-dwarf and standard-stature cultivars observed in these studies may be a result of the positive correlation between dry matter and C exudation rates from roots [84] or between aboveground dry matter and CH4 emissions [16,72,102,128]. While a reduction in emissions from semi-dwarf cultivars is oftentimes linked to reduced dry matter accumulation, Rogers et al. [93] observed a reduction in aboveground dry matter that was not accompanied by a reduction in emissions. Furthermore, Sigren et al. [130] measured greater emissions accompanied by greater soil

1

2 Figure 2. Methane emissions from standard-stature, conventional rice varieties, such as 3 "Taggart" 4 (top left) and "Wells" (top right), and hybrids varieties, such as "CLXL745" (bottom) 5 have recently been studied in the field at the Rice Research and Extension Center near **Figure 2.** Methane emissions from standard-stature, conventional rice varieties, such as "Taggart" (top left) and "Wells" (top right), and hybrids varieties, such as "CLXL745" (bottom) have recently been studied in the field at the Rice Research and Extension Center near Stuttgart, AR. Photographs taken by K. Brye.

6 Stuttgart, AR. Photographs taken by K. Brye.

acetate concentrations from a standard stature ('Mars') relative to a semi-dwarf cultivar ('Lemont'), while aboveground dry matter was similar between the two cultivars. Huang et al. [128] indicated that, while biomass may explain differences in emissions within one cultivar, the intervarietal differences in biomass are small in comparison to differences in emissions, indicating that another factor besides aboveground dry matter impacts intervarietal differen‐ ces in CH4 emissions. 44

Cultivar differences, however, extend beyond the impact of biomass production on emissions. Ma et al. [131] observed a 67% increase in CH4 oxidation from a hybrid cultivar accompanied by a reduction in emissions and soil CH4 concentration relative to pure-line cultivars. Addi‐ tional studies have also identified 25 to 37% reductions in fluxes from hybrid relative to pureline cultivars [93,132,133] (Figure 2). This indicates that greater methanotrophic activity in the rhizosphere of hybrid cultivars may reduce CH4 fluxes by oxidizing a greater proportion of the produced CH4. It is clear that cultivar selection has the potential for mitigation of CH4 from rice cultivation. However, due to the lack of understanding the mechanisms for differences in emissions, it appears that direct CH4 flux measurements from various cultivars are necessary in determining emissions differences until further research clarifies the understanding for cultivar differences in CH4 emissions (Figure 2).

#### **8. Conclusions**

Though some knowledge has been gained, there is much more that still needs to be learned and understood regarding CH4 emissions from rice production in the US, its contribution to climate change, and potential mitigation strategies. Additional field research needs to be conducted to better assess the magnitudes and relative contributions the various known factors have on CH4 production and emission from soils used for rice production.

It is possible that a single CH4 emissions factor for application to all non-California-grown, primary-crop rice in the US is too general. Consequently, the single CH4 emissions factor may be a severe overestimation for some rice-producing areas, while being an underestimation for other areas. Only after additional data have been generated can regulatory agencies, such as the USEPA, further refine greenhouse gas emissions factors to reflect the large variety of soils and agronomic cultural practices throughout the temperate US and combat the potential negative effects of climate change.

#### **Author details**

Alden D. Smartt, Kristofor R. Brye and Richard J. Norman

\*Address all correspondence to: kbrye@uark.edu

Department of Crop, Soil, and Environmental Sciences, University of Arkansas, Fayetteville, USA

#### **References**

acetate concentrations from a standard stature ('Mars') relative to a semi-dwarf cultivar ('Lemont'), while aboveground dry matter was similar between the two cultivars. Huang et al. [128] indicated that, while biomass may explain differences in emissions within one cultivar, the intervarietal differences in biomass are small in comparison to differences in emissions, indicating that another factor besides aboveground dry matter impacts intervarietal differen‐

**Figure 2.** Methane emissions from standard-stature, conventional rice varieties, such as "Taggart" (top left) and "Wells" (top right), and hybrids varieties, such as "CLXL745" (bottom) have recently been studied in the field at the

2 Figure 2. Methane emissions from standard-stature, conventional rice varieties, such as

6 Stuttgart, AR. Photographs taken by K. Brye.

Rice Research and Extension Center near Stuttgart, AR. Photographs taken by K. Brye.

4 (top left) and "Wells" (top right), and hybrids varieties, such as "CLXL745" (bottom) 5 have recently been studied in the field at the Rice Research and Extension Center near

44

Cultivar differences, however, extend beyond the impact of biomass production on emissions. Ma et al. [131] observed a 67% increase in CH4 oxidation from a hybrid cultivar accompanied

ces in CH4 emissions.

1

196 Greenhouse Gases

3 "Taggart"

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206 Greenhouse Gases

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### **Greenhouse Gases Production from Some Crops Growing Under Greenhouse Conditions**

Fernando López-Valdez , Fabián Fernández-Luqueño , Carolina Pérez-Morales and Mariana Miranda-Arámbula

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/62236

#### **Abstract**

Greenhouse gases, such as carbon dioxide (CO2), nitrous oxide (N2O) and methane (CH4), play an important role in global climate change. For example, CO2 production occurs as a result of the seasonal cycles of the biotic processes of photosynthesis and respiration, as well as through anthropogenic activities and abiotic processes such as the burning of fos‐ sil fuels. Many activities, such as Agribusiness (the production of crops and animals for food) create greenhouse gases. Our research group has studied several soil treatments such as wastewater, wastewater sludge, vermicomposting, and urea among others, in or‐ der to study the effects of soil treatments on the production of greenhouse gases (CO2, N2O, and CH4) in several cultivars, but mainly in maize, sunflower and the common bean. The principal aim of this chapter is to show how these greenhouse gases are affect‐ ed by the type of treatment, the properties of the soil, and the cultivar in question. We also look at which processes are involved in the production of CO2, N2O, and CH4 from cultivated soil. We present a review of several experiments carried out under *in vitro* or greenhouse conditions.

**Keywords:** Greenhouse gases production, wastewater sludge, fertilizers, treatments

#### **1. Introduction**

Global food demand is increasing rapidly, while the associated potential negative environ‐ mental impacts are also growing. Land clearance, the intensive use of existing croplands, inadequate agricultural management systems, and soil pollution could all contribute to an increase in the production of greenhouse gases (GHG). Understanding the future environ‐ mental impacts of global crop production, while at the same time achieving greater yields with

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

lower impacts, requires quantitative assessments of future crop demand and an understanding of how different production practices affect yields and environmental variables [1].

It is well known that crop management systems, the quality of the soil and the weathering conditions are just some of the factors used in order to assess production of GHG. Therefore, an understanding of the future environmental impacts of crop production is essential in order to achieve greater crop yields without decreasing the quality of the environment and social welfare. Additionally, Tilman [2] reported that the recent intensification of agriculture, coupled with the prospect of even further intensification in the future, will have major detrimental impacts on the world's ecosystems.

Agriculture is rightly recognized as a source of GHG production, with concomitant opportu‐ nities for its mitigation. In fact, agricultural soils can constitute either a net source or sink of the three principal GHG [3]. Soil management practices can influence GHG flux by changing at least one of the following soil properties and its associated management: 1) The soil climate (temperature and water content); 2) The physicochemical environment of the soil; 3) The soil's microorganisms (diversity and abundance); 4) The amount and chemical composition of organic or mineral fertilizers applied to the soil; and 5) Pesticides might have a strong effect on the soil microbiota (type and amount). Even a minimal change in one or more of the properties described above could control the rate and extent of GHG production and also affect the aeration and diffusion of these gases.

The objective of this chapter is to discuss how the soil production rate of GHG is affected by treatment type, soil properties, and cultivar. This review will also discuss which processes are involved in the production of CO2, N2O, and CH4 when crops are grown under *in vitro* or greenhouse conditions, and will discuss how these processes work.

#### **2. The Atmosphere, Global Climate Change and Greenhouse Gases**

The atmosphere of the Earth has evolved and changed over time and had reached a point of equilibrium. However, anthropogenic activities arising from the Industrial Revolution and subsequent development have changed Earth's atmospheric conditions. Since the industrial era began, a new phenomenon has been observed — that of global climate change (GCC). Many different sources are involved in the production of gases, with concern centering on the production of GHG in particular, as these gases are the ones implicated in the increasing rate of global warming on Earth. The main gases involved in this warming are methane (CH4), nitrous oxide (N2O), and carbon dioxide (CO2). Additional and important GHGs include water vapour, which has an effect on global climate change that can be used as a baseline with which to compare the remaining GHG. The production of these gases arises as a result of anthropo‐ genic activities, mainly the combustion of fossil fuels (CO2), extensive livestock and cattle farming (CH4), and agriculture (N2O) through denitrification or nitrification processes, and occasionally CO2 depending on the type of fertilization employed.

As we can see, global climate change is a phenomenon caused by GHG that are emitted into the atmosphere. However, the main problem is not the emission of these gases, because these gases have actually been present in the Earth's atmosphere for thousands of years and they are the products of natural processes such as volcanic eruptions, plant and animal respiration and the microbial decomposition of organic matter. The contribution of human activity has resulted in the production of large amounts of these gases and their increased concentration in the atmosphere results in global warming. The most obvious effects of global warming are the continuous increase of global temperature and the changes in atmospheric conditions. All the elements of the environment are interrelated, and as a consequence, changes in one of them lead to changes in others. Sometimes these changes are small and imperceptible, while others can be very obvious. The rate of these changes is very important because if they are too rapid, then the ability of organisms to adapt to the new conditions might not be sufficient to ensure their survival as the natural process of adaptation takes thousands and thousands of years as a part of the evolutionary processes of life on Earth. The effects of these phenomena are the extinction of species and other serious negative effects on both the agriculture and fishing industries that are important economic activities the world over.

It is important to mention that the likely impacts of global warming could be different in different types of ecosystems because of the difference in climatic conditions in those ecosys‐ tems, but the effects on the abundance and distribution of biodiversity will be constant. All of these facts suggest that the natural conditions of the planet are being seriously affected by global climate change, global warming and GHG production, so we have a serious and worldwide environmental problem to address. However, there are many strategies, such as the use of alternative sources of energy, which could be implemented around the world to mitigate the damage being caused to the planet and to promote environmental awareness with favorable results in the future.

#### **3. Experiments under** *in vitro* **conditions**

lower impacts, requires quantitative assessments of future crop demand and an understanding

It is well known that crop management systems, the quality of the soil and the weathering conditions are just some of the factors used in order to assess production of GHG. Therefore, an understanding of the future environmental impacts of crop production is essential in order to achieve greater crop yields without decreasing the quality of the environment and social welfare. Additionally, Tilman [2] reported that the recent intensification of agriculture, coupled with the prospect of even further intensification in the future, will have major

Agriculture is rightly recognized as a source of GHG production, with concomitant opportu‐ nities for its mitigation. In fact, agricultural soils can constitute either a net source or sink of the three principal GHG [3]. Soil management practices can influence GHG flux by changing at least one of the following soil properties and its associated management: 1) The soil climate (temperature and water content); 2) The physicochemical environment of the soil; 3) The soil's microorganisms (diversity and abundance); 4) The amount and chemical composition of organic or mineral fertilizers applied to the soil; and 5) Pesticides might have a strong effect on the soil microbiota (type and amount). Even a minimal change in one or more of the properties described above could control the rate and extent of GHG production and also affect

The objective of this chapter is to discuss how the soil production rate of GHG is affected by treatment type, soil properties, and cultivar. This review will also discuss which processes are involved in the production of CO2, N2O, and CH4 when crops are grown under *in vitro* or

The atmosphere of the Earth has evolved and changed over time and had reached a point of equilibrium. However, anthropogenic activities arising from the Industrial Revolution and subsequent development have changed Earth's atmospheric conditions. Since the industrial era began, a new phenomenon has been observed — that of global climate change (GCC). Many different sources are involved in the production of gases, with concern centering on the production of GHG in particular, as these gases are the ones implicated in the increasing rate of global warming on Earth. The main gases involved in this warming are methane (CH4), nitrous oxide (N2O), and carbon dioxide (CO2). Additional and important GHGs include water vapour, which has an effect on global climate change that can be used as a baseline with which to compare the remaining GHG. The production of these gases arises as a result of anthropo‐ genic activities, mainly the combustion of fossil fuels (CO2), extensive livestock and cattle farming (CH4), and agriculture (N2O) through denitrification or nitrification processes, and

As we can see, global climate change is a phenomenon caused by GHG that are emitted into the atmosphere. However, the main problem is not the emission of these gases, because these

**2. The Atmosphere, Global Climate Change and Greenhouse Gases**

greenhouse conditions, and will discuss how these processes work.

occasionally CO2 depending on the type of fertilization employed.

of how different production practices affect yields and environmental variables [1].

detrimental impacts on the world's ecosystems.

210 Greenhouse Gases

the aeration and diffusion of these gases.

We are interested in understanding which soil processes are involved in GHG production, and how they work, in several treatments or fertilizers. The production of CO2, N2O, and CH4 when crops were grown under *in vitro* or greenhouse conditions was studied. These experiments were carried out using different types of soil (nitrogen depleted and/or alkaline-saline) and several crops were studied.

One of the experiments was conducted in order to investigate the evolution of nitrogen and its loss as a part of the nitrogen cycle. Different fertilizers or treatments were tested. These were ammonium sulphate [(NH4)2SO4, 200 mg NH4 <sup>+</sup> kg-1], wastewater sludge (200 mg NH4 <sup>+</sup> kg-1), sterile wastewater sludge (200 mg NH4 <sup>+</sup> kg-1) and a control (distilled H2O). All of the treatments were added with KNO3 at 100 mg N kg-1, in two different soils (one agricultural and N depleted soil, and the second a saline-alkaline and N depleted soil) at 40% of water holding capacity (WHC) under *in vitro* conditions for 56 days. The variables were CO2, N2O, NH4 + -N, NO2 - -N, and NO3 - -N, and were measured and assessed [4]. The soils used were specifics, i.e. the first soil was an agricultural soil which had mainly been cultivated with maize (> 25 years), low fertilization (< 50 kg N ha-1) and was both C and N depleted (6.5 g organic C kg-1, 0.2 g total N Kjeldahl kg-1, pH 7.8, electrolytic conductivity (EC) 1.0 dS m-1, and the textural soil classification was loamy sand) from Otumba, in the State of Mexico (Mexico) (19° 42' N, 98° 49' W). The second soil was classified as an uncultivated soil (some grasses and small trees could be found) as a result of its former lake bed origin. It was found to be N depleted and alkaline-saline, pH 10.3, EC 12.4 dS m-1, 49 g organic C kg-1, and 0.6 total N Kjeldahl kg-1, from Texcoco, State of Mexico (Mexico) (19° 30' N, 98° 53' W). The results showed that production of CO2 from the Otumba soil was not affected by the addition of NH4 + or NO3 - , i.e. both fertilizers produced a similar amount of CO2, approximately 350 mg CO2-C kg-1 dry soil. The sterilized sewage sludge increased the production of CO2, > 1,000 mg CO2-C kg-1 dry soil, i.e. over twice the production compared with that of the controls (soils treated with ammonium or nitrate). When wastewater sludge was added, the CO2 production was ~ 3,100 mg CO2-C kg-1 dry soil, a value twice that of sterilized sludge and eight times that of the controls. In the Texcoco soil, a similar contour was found in the CO2-C dynamics. The control treatments present > 350 mg CO2-C kg-1 dry soil. The soil treated with sterilized sludge showed a concentration of CO2 that was > 1,100 mg CO2-C kg-1 dry soil (2.1 times that of the controls) while the treatment with sewage sludge was ~ 2,100 mg CO2-C kg-1 dry soil (over five times that of the controls). Different soils showed similar contours in C dynamics, and these results revealed that CO2 production was not particularly different and that the processes involved were rather similar in the different soils investigated. Wastewater sludge is characterized by large amounts of organic matter (organic C) and it's suppose got a large amounts of organic matter easily decomposable. So, the sludge is easily and rapidly mineralized in both soils [4]. When sewage sludge was sterilized, the microbiota were destroyed and this might have affected the properties of the organic matter in the sludge [4]. It could be that the organic matter was more readily accessible for the soil microorganisms, so that the production of CO2 should be higher in soils with sterilized sludge, but the results revealed that this is not necessarily true. The results also showed that the soil microorganisms and the sludge microorganisms could be exerting a synergistic action on the degradation of organic matter because the degradative action of the microorganisms of the soil or the sludge alone cannot improve the degradation of organic matter in the treated soil or in the sludge [4].

The ammonium dynamics showed that the initial concentrations of N were reduced after the first 3 days, and after that, a release of the mineral occurred from day 3 up to day 14. Later still, the concentration of ammonium decreased by up to < 14 mg N kg-1 dry soil for all the treatments in both the Otumba and Texcoco soils, and the ammonium concentration decreased by up to < 2 mg N kg-1 dry soil for all treatments, except for the soil treated with sterilized sludge, < 31 mg N kg-1 dry soil. The contour of the ammonium dynamics was similar in both the Otumba and Texcoco soils. Many abiotic and biotic processes might affect the concentration of NH4 + in soil, such as NH4 + fixation in the soil matrix, volatilisation of NH3, and immobilization or oxidation of NH4 + . Some soil processes were occurring at too low a level to be detectable, such as NH4 + fixation and the volatilisation of NH3. The nitrate dynamics were similar in both soils. The concentration of NO3 - was ~120 mg N kg-1 dry soil in the control treatment in both soils. The ammonium concentration was similar in both soils, > 200 mg N kg-1 dry soil, treatments with sludge reached > 255 mg N kg-1 dry soil and > 300 mg N kg-1 dry soil in the Texcoco and Otumba soils respectively, and soils treated with sterilized sludge increased the concentration to > 300 mg N kg-1 dry soil in the Texcoco soil, while in the Otumba soil it was > 325 mg N kg-1 dry soil. These results suggest that soil fertilized with wastewater sludge showed an increased NO3 concentration with a hypothetic mineralisation of ~ 60% at day 56 [4].

Kjeldahl kg-1, pH 7.8, electrolytic conductivity (EC) 1.0 dS m-1, and the textural soil classification was loamy sand) from Otumba, in the State of Mexico (Mexico) (19° 42' N, 98° 49' W). The second soil was classified as an uncultivated soil (some grasses and small trees could be found) as a result of its former lake bed origin. It was found to be N depleted and alkaline-saline, pH 10.3, EC 12.4 dS m-1, 49 g organic C kg-1, and 0.6 total N Kjeldahl kg-1, from Texcoco, State of Mexico (Mexico) (19° 30' N, 98° 53' W). The results showed that production of CO2 from the

similar amount of CO2, approximately 350 mg CO2-C kg-1 dry soil. The sterilized sewage sludge increased the production of CO2, > 1,000 mg CO2-C kg-1 dry soil, i.e. over twice the production compared with that of the controls (soils treated with ammonium or nitrate). When wastewater sludge was added, the CO2 production was ~ 3,100 mg CO2-C kg-1 dry soil, a value twice that of sterilized sludge and eight times that of the controls. In the Texcoco soil, a similar contour was found in the CO2-C dynamics. The control treatments present > 350 mg CO2-C kg-1 dry soil. The soil treated with sterilized sludge showed a concentration of CO2 that was > 1,100 mg CO2-C kg-1 dry soil (2.1 times that of the controls) while the treatment with sewage sludge was ~ 2,100 mg CO2-C kg-1 dry soil (over five times that of the controls). Different soils showed similar contours in C dynamics, and these results revealed that CO2 production was not particularly different and that the processes involved were rather similar in the different soils investigated. Wastewater sludge is characterized by large amounts of organic matter (organic C) and it's suppose got a large amounts of organic matter easily decomposable. So, the sludge is easily and rapidly mineralized in both soils [4]. When sewage sludge was sterilized, the microbiota were destroyed and this might have affected the properties of the organic matter in the sludge [4]. It could be that the organic matter was more readily accessible for the soil microorganisms, so that the production of CO2 should be higher in soils with sterilized sludge, but the results revealed that this is not necessarily true. The results also showed that the soil microorganisms and the sludge microorganisms could be exerting a synergistic action on the degradation of organic matter because the degradative action of the microorganisms of the soil or the sludge alone cannot improve the degradation of organic matter in the treated soil

The ammonium dynamics showed that the initial concentrations of N were reduced after the first 3 days, and after that, a release of the mineral occurred from day 3 up to day 14. Later still, the concentration of ammonium decreased by up to < 14 mg N kg-1 dry soil for all the treatments in both the Otumba and Texcoco soils, and the ammonium concentration decreased by up to < 2 mg N kg-1 dry soil for all treatments, except for the soil treated with sterilized sludge, < 31 mg N kg-1 dry soil. The contour of the ammonium dynamics was similar in both the Otumba and Texcoco soils. Many abiotic and biotic processes might affect the concentration of NH4

fixation in the soil matrix, volatilisation of NH3, and immobilization or

. Some soil processes were occurring at too low a level to be detectable, such


fixation and the volatilisation of NH3. The nitrate dynamics were similar in both soils.

The ammonium concentration was similar in both soils, > 200 mg N kg-1 dry soil, treatments with sludge reached > 255 mg N kg-1 dry soil and > 300 mg N kg-1 dry soil in the Texcoco and Otumba soils respectively, and soils treated with sterilized sludge increased the concentration

+ or NO3 -

, i.e. both fertilizers produced a

+ in

Otumba soil was not affected by the addition of NH4

or in the sludge [4].

212 Greenhouse Gases

soil, such as NH4

oxidation of NH4

The concentration of NO3

as NH4 + +

+

The production of N2O was measured in both soils for seven days under C2H2 (10% v/v) conditions. The control treatment showed a N2O production of < 0.02 mg N kg-1 with or without C2H2 in the Otumba soil. N2O production increased when ammonium was added to the Otumba soil, 0.04 mg N kg-1, but the addition of C2H2 reduced it to 0.01 mg N kg-1. When the sterilized and non-sterilized sludge were added to the Otumba soil, N2O production increased to 1 mg N kg-1 dry soil without C2H2 and 0.49 mg N kg-1 dry soil with C2H2. In the Texcoco soil, the control treatment was below 0.005 mg N kg-1 dry soil, but when C2H2 was added an increase to 0.09 mg N kg-1 dry soil was observed. Soil treated with ammonium increased the production of N2O (0.04 mg N kg-1), but under C2H2 conditions the concentration was low at 0.01 mg N kg-1 dry soil when compared with the control treatment. Soils treated with sterilized and nonsterilized sludge increased the production of N2O (2.1 mg N kg-1 and 0.75 mg N kg-1) compared with the control treatment, however, the addition of C2H2 increased the N2O concentration of soil treated with sludge (2.1 mg N kg-1 dry soil) compared with soil treated with sterilized sludge (1.8 mg N kg-1). It could be argued that the addition of easily decomposable organic matter into the soil will induce the denitrification process, where NO3 is reduced to N2O and N2 as final products. There are factors that could be important in controlling the production of N2O in denitrification, such as oxygen, pH and the ratio of nitrate / available carbon [5]. There are additional parallel factors for NO that are less well understood [5]. It was evident that in a soil treatment of wastewater sludge or sterilized sludge, the N loss was increased. The microorganisms from the soil and the sludge were found to be working together in order to degrade the organic matter in the sludge. In addition, it could be suggested that more denitrifiers may be present in the Texcoco soil than in the Otumba soil, and a decrease of 50 mg NO3 - -N was counted at day 56, and the N2O concentration was approximately double (plus 1 mg N2O kg-1 dry soil) that of the Otumba soil on the day 7 [4]. When NO3 - (an e acceptor) is present in excess compared to organic C (an e donor), the denitrifiers could be said to be "spendthrift" with respect to NO3 and in general produce N2O as the major product. When the same NO3 - is limited, the denitrifiers use it to its maximum potential as an e acceptor and reduce it all to N2 (dinitrogen) [5]. Schimel and Holland (2005) explain that while the major producer of NO is nitrification, N2O can also be produced in large quantities by nitrification or denitrification due to the fact that it is less reactive and can outflow from soils — even wet soils.

In our study, the microorganisms in the soil and sludge acted together synergistically in the reduction process, i.e. N2O to N2, the denitrification process, and in the Texcoco soil under C2H2 conditions, but the main contribution was from the soil microorganisms under the same conditions. In particular, the production of N2 was almost 50% of the total gas evolved. When sterilized sludge was applied to the Texcoco soil, minimal N2 was produced (~0.2 mg N2O-N kg-1). Furthermore, untreated soil showed an increase in N2O when C2H2 was added (~0.07 mg N2O-N kg-1) and the N2 produced was approximately 75% of the total gas evolved, under the conditions established. In the Otumba soil, the N2O was produced by nitrification process (at 40% WHC), and showed in the soil treated with ammonium, sludge and sterilized sludge at 0.04, 1.0, and 1.0 mg N2O-N kg-1 respectively, when compared with untreated soil. N2O production was low when C2H2 was added to the Otumba soil, and no N2 was produced under these conditions. It can be established that when organic matter with a high N content is added to soil, it significantly increases N2O production compared to untreated soil or soil fertilized with ammonium [(NH4)2SO4] in both soils (an ordinary soil and an alkaline-saline soil). The major source of production of N2O was found in all treated soils to be as a result of the nitrification process, and the production of N2 was not recorded in the Otumba soil in this experiment. In the Texcoco soil, the major source of production of N2O was as a result of the denitrification process by microorganisms in the soil, and the production of N2 was approxi‐ mately 50% of the total gas evolved (1.1 mg N2O-N kg-1).

#### **4. Experiments under greenhouse conditions**

#### **4.1. First experiment under greenhouse conditions**

Subsequent studies were established in order to better understand plant growth and the production of GHG (CO2 and N2O) when a regular *Bacillus subtilis* strain was inoculated on the surface of the sunflower (*Helianthus annuus* L.) cultivar seeds under greenhouse conditions. The *B. subtilis* strain was characterized as PGPR, i.e. showing antagonistic activity against *Fusarium oxysporum* and *Rhizoctonia solani* AG1, phosphate solubilizing activity, 1-aminocy‐ clopropane-1-carboxilate deaminase, and indole-3-acetic acid production. The strain was found as regular PGPR, for more details see [6].

The soil was collected from Alcholoya (Acatlán, in the State of Hidalgo, Mexico). This soil is an agricultural soil with a pH of 6.5, electrolytic conductivity (EC) 0.7 dS m-1, 846 g kg-1, organic C content was 11.1 g C kg-1 soil, and total N content 1.0 g N kg-1 soil. The soil was sampled from three different plots (400 m2 ), ~ 800 kg was obtained and each plot was pooled separately and passed through a 5 mm sieve. Thirty-six sub-samples of 6.5 kg of soil from each plot were placed in cylindrical pots (Ø = 16 cm, 50 cm) with 7 cm of gravel in the bottom. Four treatments were applied, with nine pots for each of the three soil sites sampled (*n* = 27). The first treatment was unfertilized and uncultivated soil (used as the CONTROL treatment), the second treat‐ ment was unfertilized soil cultivated with sunflowers (SUNFLOWER treatment), the third treatment was soil cultivated with sunflowers and fertilized with 0.5 g urea (75 kg N ha-1, UREA treatment), and the fourth treatment was soil cultivated with sunflowers (seeds were dressed with *B. subtilis*) and fertilized with 0.5 g urea (BS treatment). All treatments were irrigated with tap water, with an additional input of 19 kg mineral-N ha-1 as NO2 - -N and NO3 - -N. In addition, twelve days after the emergence of the plantlets, they were fertilized with another 0.5 g of urea (the UREA and BS treatments), giving a total amount of 150 kg N ha-1; and three weeks after sowing the plantlets were drenched with 4 mL of a bacterial suspension (at the same concen‐ tration as described above) adjacent to the plantlet roots at a depth of 3 cm. From the beginning of the experiments and approximately every two days for the following 30 days, the pots were closed air-tight and their atmosphere was analysed for CO2 and N2O at times of 0, 3, 15 and 30 mins. The experiment was repeated twice [6].

40% WHC), and showed in the soil treated with ammonium, sludge and sterilized sludge at 0.04, 1.0, and 1.0 mg N2O-N kg-1 respectively, when compared with untreated soil. N2O production was low when C2H2 was added to the Otumba soil, and no N2 was produced under these conditions. It can be established that when organic matter with a high N content is added to soil, it significantly increases N2O production compared to untreated soil or soil fertilized with ammonium [(NH4)2SO4] in both soils (an ordinary soil and an alkaline-saline soil). The major source of production of N2O was found in all treated soils to be as a result of the nitrification process, and the production of N2 was not recorded in the Otumba soil in this experiment. In the Texcoco soil, the major source of production of N2O was as a result of the denitrification process by microorganisms in the soil, and the production of N2 was approxi‐

Subsequent studies were established in order to better understand plant growth and the production of GHG (CO2 and N2O) when a regular *Bacillus subtilis* strain was inoculated on the surface of the sunflower (*Helianthus annuus* L.) cultivar seeds under greenhouse conditions. The *B. subtilis* strain was characterized as PGPR, i.e. showing antagonistic activity against *Fusarium oxysporum* and *Rhizoctonia solani* AG1, phosphate solubilizing activity, 1-aminocy‐ clopropane-1-carboxilate deaminase, and indole-3-acetic acid production. The strain was

The soil was collected from Alcholoya (Acatlán, in the State of Hidalgo, Mexico). This soil is an agricultural soil with a pH of 6.5, electrolytic conductivity (EC) 0.7 dS m-1, 846 g kg-1, organic C content was 11.1 g C kg-1 soil, and total N content 1.0 g N kg-1 soil. The soil was sampled

and passed through a 5 mm sieve. Thirty-six sub-samples of 6.5 kg of soil from each plot were placed in cylindrical pots (Ø = 16 cm, 50 cm) with 7 cm of gravel in the bottom. Four treatments were applied, with nine pots for each of the three soil sites sampled (*n* = 27). The first treatment was unfertilized and uncultivated soil (used as the CONTROL treatment), the second treat‐ ment was unfertilized soil cultivated with sunflowers (SUNFLOWER treatment), the third treatment was soil cultivated with sunflowers and fertilized with 0.5 g urea (75 kg N ha-1, UREA treatment), and the fourth treatment was soil cultivated with sunflowers (seeds were dressed with *B. subtilis*) and fertilized with 0.5 g urea (BS treatment). All treatments were irrigated with

twelve days after the emergence of the plantlets, they were fertilized with another 0.5 g of urea (the UREA and BS treatments), giving a total amount of 150 kg N ha-1; and three weeks after sowing the plantlets were drenched with 4 mL of a bacterial suspension (at the same concen‐ tration as described above) adjacent to the plantlet roots at a depth of 3 cm. From the beginning of the experiments and approximately every two days for the following 30 days, the pots were

), ~ 800 kg was obtained and each plot was pooled separately





mately 50% of the total gas evolved (1.1 mg N2O-N kg-1).

**4. Experiments under greenhouse conditions**

**4.1. First experiment under greenhouse conditions**

found as regular PGPR, for more details see [6].

tap water, with an additional input of 19 kg mineral-N ha-1 as NO2

from three different plots (400 m2

214 Greenhouse Gases

The daily CO2 production rate for some treatments was large at the beginning of the experi‐ ments (data not shown). The daily CO2 production rate showed a drop, remaining < 5 mg C kg-1 day-1 on day 2 and after that it remained at < 8 mg C kg-1 day-1 for all treatments up to end of the experiments. The mean CO2 production rate was not significantly different between treatments. The daily N2O production rate remained ≤ 0.75 µg N kg-1 day-1 for the SUNFLOW‐ ER and CONTROL treatments. Meanwhile, in the BS and UREA treatments, the daily N2O production rate remained ≤ 2.1 µg N kg-1 day-1 with a maximum score of production in the first 14 days. The mean of the N2O production rate of BS treatment was significantly high when compared with the SUNFLOWER and CONTROL treatments. Cultivating soil with sunflow‐ ers (SUNFLOWER treatment) did not affect the production of CO2 compared with the CONTROL treatment (uncultivated soil). It is well known that cultivated soil frequently increases the production of CO2, possibly due to the activities of the microorganisms degrading the easily decomposable organic matter such as the dying roots and root exudates in the rhizosphere, thereby increasing the production of CO2 from the soil. Soils cultivated with sunflowers and fertilized with urea did not affect the production of CO2. Applying urea to soil commonly has no effect on the production of CO2 from soils. However, the production of CO2 might be stimulated when urea is applied to N depleted soil as reported in Phillips and Podrebarac (2009), where the CO2 production was tripled when 112 kg urea-N kg-1 was applied to an arable soil. Increases in CO2 production with several doses of urea-N application indicate that agronomic-scale N inputs might stimulate microbial carbon cycling in arable soils [7]. The inoculation of *B. subtilis* on sunflower roots in soils fertilized with urea did not show any effect on the production of CO2 compared with the UREA treatment. The application of UREA to the soil resulted in more than a doubling of the mean N2O production rate when compared with the CONTROL treatment. The production of N2O and NO in the soil is the result of factors such as ion concentrations and soil conditions under chemical disequilibrium, i.e. the oxidative process of NH4 + to NO3 under aerobic conditions, nitrification, and a reductive process of NO3 to N2 under anaerobic conditions, and denitrification or the nitrifier denitrification [5, 8, 9]. According to the IPCC, N2O is the main gas produced and released to the atmosphere by soil microorganisms [9]. So, when urea is applied into the soil, hydrolysis of the urea is immediately started, releasing NH4 <sup>+</sup> into the soil, which is rapidly transformed to NO3 with the simultaneous production of N2O by the same process. A high concentration of NO3 in soil favours the production of N2O due to the presence of anaerobic micro-sites in highly com‐ pacted soil.

A principal component analysis (PCA) was undertaken to investigate several plant biometric parameters and soil properties in the production of N2O and CO2. The analysis was carried out to include all variables. The PCA revealed that BS treatment has an effect on the shoots of plants, i.e. the shoot length, dry weight of shoot, and the fresh weight of shoot. UREA treatment has an effect on roots, i.e. the dry weight of roots, length of roots, and fresh weight of roots, and a minimal effect on seed weight. Otherwise, the soil properties PCA showed that both the BS and UREA treatments had an effect on NO3 and EC at depths of 0-15 cm and 15-30 cm. The BS treatment also has effect on the production of N2O and the UREA treatment has an effect on NH4 <sup>+</sup> at a depth 0-15 cm and a slight effect on CO2 production. These results were found to correspond with a production rate of N2O that was ≤ 2.1 µg N kg-1 day-1 with a maximum production in the first 14 days of more than that in the UREA treatment (< 1.5 µg N kg-1 day-1). Additionally, the PCA showed that *B. subtilis*, as a regular strain, had a marked effect on the production of N2O but not on CO2 production. This strain might be involved in nitrifier denitrification (or aerobic nitrification-denitrification) as was reported by Kim et al. [10] and Yang et al. [11]. Both research groups demonstrated that the *Bacillus* genus is involved in nitrification and denitrification, namely *B. subtilis*, *B. cereus* and *B. licheniformis*, where *B. subtilis* is involved in the nitrification process and *B. licheniformis* is involved in denitrification or aerobic denitrification [10]. Yang et al. reported that the strain of *Bacillus subtilis* A1 is an aerobic heterotrophic nitrifying–denitrifying bacterium, which is able to convert NH4 + to N2 under fully aerobic conditions, while growing either autotrophically or heterotrophically [11]. In our experiment, the environmental condition of the soil was aerobic throughout the experiment, so it could be hypothesized that the *B. subtilis* strain could be involved in the nitrification process or the nitrifier denitrification process (or aerobic denitrification).

#### **4.2. Second experiment under greenhouse conditions**

The second experiment was carried out using wastewater sludge as an organic fertilizer. It tested the effect of the sludge or urea on sunflower growth, and the effect of some soil properties on the production of CO2 and N2O. The plant characteristics were also evaluated. Wastewater sludge or sewage sludge is generated during wastewater treatment and is an unavoidable by-product. However, the sludge can be seen as an invaluable by-product when it is applied to the soil as stabilized sludge. Also, this waste management is the most economical form of disposal employed to reduce the large amount of sewage sludge. In addition, waste‐ water sludge is organic matter rich in minerals and is an outstanding source of C and N, *inter alia*. Applying sludge to the soil offers the opportunity of recycling nutrients for use by plants, while at the same time returning C as organic matter to the soil in order to improve agriculture processes.

This experiment was carried out in the same way as the previous experiment (described above). The wastewater sludge was collected from the Reciclagua treatment plant, S.A. de C.V., where wastewater from various industries (including the food industry) and households is treated. The properties of the sewage sludge were pH 8.1, EC 7.9 dS m-1, water content 847 g kg-1, organic C content 288 g kg-1, total N content 41.8 g kg-1, NH4 + 13 g N kg-1, NO2 - 8.3 mg N kg-1, and NO3 - 122 mg N kg-1. For more details of these, see [12]. Four treatments were established in cylindrical pots, comprising nine soil samples from three sampled plots (*n* = 27). The treatments were: i) unfertilized and unsown soil (CONTROL treatment), ii) unfertilized soil cultivated with sunflowers (SUNFLOWER treatment), iii) cultivated soil fertilized with 0.5 g urea (0.5 g urea × 2 applications, equivalent to 150 kg N ha-1, UREA treatment), and iv) soil cultivated and fertilized with 30 g sludge (SLUDGE treatment). The sludge was added so as to be equivalent to 150 kg N ha-1, assuming that sludge mineralisation was 40% mineral N during the crop cycle. Tap water supplied a total amount of 19 kg mineral-N ha-1 through irrigation in all treatments and throughout the experiment. In order to measure the production of gases, the pots were closed airtight approximately every two days for the first 30 days, and their atmospheres were analysed for CO2 and N2O at sequential times of 0, 3, 15 and 30 mins. The experiment was replicated twice [12].

BS treatment also has effect on the production of N2O and the UREA treatment has an effect

correspond with a production rate of N2O that was ≤ 2.1 µg N kg-1 day-1 with a maximum production in the first 14 days of more than that in the UREA treatment (< 1.5 µg N kg-1 day-1). Additionally, the PCA showed that *B. subtilis*, as a regular strain, had a marked effect on the production of N2O but not on CO2 production. This strain might be involved in nitrifier denitrification (or aerobic nitrification-denitrification) as was reported by Kim et al. [10] and Yang et al. [11]. Both research groups demonstrated that the *Bacillus* genus is involved in nitrification and denitrification, namely *B. subtilis*, *B. cereus* and *B. licheniformis*, where *B. subtilis* is involved in the nitrification process and *B. licheniformis* is involved in denitrification or aerobic denitrification [10]. Yang et al. reported that the strain of *Bacillus subtilis* A1 is an aerobic heterotrophic nitrifying–denitrifying bacterium, which is able to convert NH4

under fully aerobic conditions, while growing either autotrophically or heterotrophically [11]. In our experiment, the environmental condition of the soil was aerobic throughout the experiment, so it could be hypothesized that the *B. subtilis* strain could be involved in the

The second experiment was carried out using wastewater sludge as an organic fertilizer. It tested the effect of the sludge or urea on sunflower growth, and the effect of some soil properties on the production of CO2 and N2O. The plant characteristics were also evaluated. Wastewater sludge or sewage sludge is generated during wastewater treatment and is an unavoidable by-product. However, the sludge can be seen as an invaluable by-product when it is applied to the soil as stabilized sludge. Also, this waste management is the most economical form of disposal employed to reduce the large amount of sewage sludge. In addition, waste‐ water sludge is organic matter rich in minerals and is an outstanding source of C and N, *inter alia*. Applying sludge to the soil offers the opportunity of recycling nutrients for use by plants, while at the same time returning C as organic matter to the soil in order to improve agriculture

This experiment was carried out in the same way as the previous experiment (described above). The wastewater sludge was collected from the Reciclagua treatment plant, S.A. de C.V., where wastewater from various industries (including the food industry) and households is treated. The properties of the sewage sludge were pH 8.1, EC 7.9 dS m-1, water content 847 g kg-1, organic

 122 mg N kg-1. For more details of these, see [12]. Four treatments were established in cylindrical pots, comprising nine soil samples from three sampled plots (*n* = 27). The treatments were: i) unfertilized and unsown soil (CONTROL treatment), ii) unfertilized soil cultivated with sunflowers (SUNFLOWER treatment), iii) cultivated soil fertilized with 0.5 g urea (0.5 g urea × 2 applications, equivalent to 150 kg N ha-1, UREA treatment), and iv) soil cultivated and fertilized with 30 g sludge (SLUDGE treatment). The sludge was added so as to be equivalent to 150 kg N ha-1, assuming that sludge mineralisation was 40% mineral N during the crop cycle. Tap water supplied a total amount of 19 kg mineral-N ha-1 through irrigation in all treatments

+

13 g N kg-1, NO2

nitrification process or the nitrifier denitrification process (or aerobic denitrification).

**4.2. Second experiment under greenhouse conditions**

C content 288 g kg-1, total N content 41.8 g kg-1, NH4

<sup>+</sup> at a depth 0-15 cm and a slight effect on CO2 production. These results were found to

+ to N2


on NH4

216 Greenhouse Gases

processes.

NO3 - The results showed that the CONTROL, SUNFLOWER and UREA treatments were not significantly different with respect to the production rate of CO2 (1.59, 2.03 and 2.6 mg C kg-1 day-1, respectively) (Table 1). The CO2 production rate of sunflowers cultivated in soil fertilized with sewage sludge (SLUDGE treatment) was 2.96 mg C kg-1 day-1 and was significantly different compared with the CONTROL treatment. In other words, soil cultivated and fertilized with sewage sludge was equivalent to both the soil cultivated and fertilized with urea, and the unfertilized soil. It should be taken in account that several factors affect CO2 production in the soil, such as rhizosphere respiration and soil microbial respiration, soil moisture, soil temperature, substrate quantity and quality, vegetation type, and land use and management regimes [13].


**Table 1.** The production rates of CO2 and N2O from soil cultivated with *H. annuus* under greenhouse conditions.

On the other hand, the rate of production N2O was significantly different in cultivated soil fertilized with sludge, 2.5 µg N kg-1 day-1 compared with the remaining treatments, 0.7, 0.3, and 0.2 µg N kg-1 day-1 for the UREA, SUNFLOWER, and CONTROL treatments respectively. As previously discussed, wastewater sludge is an organic matter which is rich in easily decomposable material. In addition, it has been demonstrated that microorganisms from the sludge plus those from the soil work together synergistically to accelerate the decomposition of organic matter [6]. The high levels of microbial activity stimulated by the addition of material with high C and N contents could increase the production of both CO2 and N2O. According to Kool et al. [9], the loss of N in our experiment might primarily be as a result of the nitrifier nitrification process (from ammonium oxidation) followed by the nitrifier denitrification process. As a result, the wastewater sludge had a NH4 + concentration of 13 g N kg-1, an amount of ammonium which might not be oxidized to N2O so rapidly. The chemical composition of organic or mineral fertilizers — or even residues applied to the soil — is an important factor in regulating the magnitude of N2O production.

A PCA was performed on all relevant properties of the soil. The first principal component explained about 22% of the observed variation, while the second accounted for 17% of the observed variation. On the related scatter plot, the UREA treatment lies in upper right quadrant and the SLUDGE treatment is in the upper left quadrant. The SUNFLOWER and CONTROL treatments were found in lower left and right quadrants respectively (Figure 1).

**Figure 1.** Results of a principal component analysis (PCA) performed on soil properties under greenhouse conditions (*n* = 27).

The UREA treatment seems to have an effect on the pH of soil at both depths (0–15 and 15–30 cm) and also affects the production of CO2 and N2O. The PCA also revealed that the SLUDGE treatment affected EC at a depth of 0–15 cm and had a slight effect on both ammonium and nitrate concentrations at both depths.

#### **4.3. Third experiment under greenhouse conditions**

In this study, the effect of urea, wastewater sludge and vermicomposting on the production of CO2 and N2O was investigated. The Otumba soil (State of Mexico, Mexico), which was characterized as a sandy loam with pH 7.6, EC 1.15 dS m-1, and an organic C content of 7.2 g C kg-1, was used in this study. Wastewater sludge was again collected from Reciclagua S.A. de C.V. (as described above). The vermicompost was prepared with wastewater sludge from Reciclagua and *Eisenia fetida*. The vermicompost was obtained from a mixture of sludge (1,800 g) and manure (800 g) with 70% water content, and was added to 40 individuals of *E. fetida* with the mixture being conditioned over three months. The properties of the vermicompost were pH 7.9, EC 11 dS m-1, organic C content 163 g kg-1, and a total N content of 2 g kg-1. As

well, the vermicompost presented < 3 CFU g-1 *Salmonella* sp., no *Shigella* sp., and no helminth ova.

A PCA was performed on all relevant properties of the soil. The first principal component explained about 22% of the observed variation, while the second accounted for 17% of the observed variation. On the related scatter plot, the UREA treatment lies in upper right quadrant and the SLUDGE treatment is in the upper left quadrant. The SUNFLOWER and CONTROL treatments were found in lower left and right quadrants respectively (Figure 1).

SLUDGE

UREA

pH, 15-30 cm pH, 0-15 cm

CONTROL

*Factor 2: 17%*

, 15-30 cm Production of CO2

EC, 0-15 cm EC, 15-30 cm

+, 15-30 cm Production of N2O


**Figure 1.** Results of a principal component analysis (PCA) performed on soil properties under greenhouse conditions

The UREA treatment seems to have an effect on the pH of soil at both depths (0–15 and 15–30 cm) and also affects the production of CO2 and N2O. The PCA also revealed that the SLUDGE treatment affected EC at a depth of 0–15 cm and had a slight effect on both ammonium and

In this study, the effect of urea, wastewater sludge and vermicomposting on the production of CO2 and N2O was investigated. The Otumba soil (State of Mexico, Mexico), which was characterized as a sandy loam with pH 7.6, EC 1.15 dS m-1, and an organic C content of 7.2 g C kg-1, was used in this study. Wastewater sludge was again collected from Reciclagua S.A. de C.V. (as described above). The vermicompost was prepared with wastewater sludge from Reciclagua and *Eisenia fetida*. The vermicompost was obtained from a mixture of sludge (1,800 g) and manure (800 g) with 70% water content, and was added to 40 individuals of *E. fetida* with the mixture being conditioned over three months. The properties of the vermicompost were pH 7.9, EC 11 dS m-1, organic C content 163 g kg-1, and a total N content of 2 g kg-1. As

NO2 - , 0-15 cm

SUNFLOWER


(*n* = 27).

218 Greenhouse Gases

*Factor 1: 22%*

nitrate concentrations at both depths.

**4.3. Third experiment under greenhouse conditions**

NO3 - , 15-30 cm

NH4 +, 0-15 cm

NH4

NO2 -

NO3 - , 0-15 cm Forty-five sub-samples of 3.25 kg soil were prepared, i.e. three soil samples, three replicates, and five treatments were established. The treatments were: i) soil fertilized with 0.07 g urea kg-1, (UREA treatment), ii) soil fertilized with 21.2 g sewage sludge kg-1 (H-SLUDGE treat‐ ment), iii) soil fertilized with 12.8 g sewage sludge kg-1 (L-SLUDGE treatment), iv) soil fertilized with 81.5 g vermicompost kg-1 (VERMI treatment), and v) unfertilized soil (CONTROL treatment). The amount of urea was 80 kg N ha-1 for the UREA, H-SLUDGE, and VERMI treatments, while the L-SLUDGE treatment was 48 kg N ha-1. The production of CO2 and N2O was analysed every two days at 0, 3, 15 and 30 mins, until day 97. The cultivar of common bean used was Negro-8025 (from Universidad Autónoma Chapingo, Texcoco, State of Mexico). Tap water was used for irrigation at a rate of 500 mL every seven days. The experiments were triplicated and ran for 117 days in total.

A large amount of CO2 was recorded at the end the experiment (from day 45 until day 82). The UREA treatment had no significant effect on CO2 production when compared with the CONTROL treatment (0.043 mg C kg-1 dry soil). Wastewater sludge was found to increase the mean production of CO2 in the soil at 0.064 mg C kg-1 compared with the untreated soil, and the VERMI treatment showed the largest mean CO2 production at 0.1 mg C kg-1. Urea had no significant effect on CO2 production when compared with the CONTROL treatment. Occa‐ sionally, urea might stimulate the production of CO2 in N depleted soils. Wastewater sludge increased CO2 production through the mineralization of organic C and the increasing of microbial activity in the rhizosphere. Similarly, vermicomposting stimulated plant growth and increased root exudates and microbial activity in the rhizosphere. The high levels of CO2 produced towards the end of the experiments could be related to rapid root growth and the beginning of root decomposition. For example, common bean plants at 49 days after sowing (flowering and nodule senescence begins) showed a decrease in the number of nodules, and their nodule cell walls slowly became thinner and degraded [14] thereby improving the environment for the growth of microorganisms and increasing the production of CO2 towards the end of the experiments.

The mean production of N2O was -0.004 µg N kg-1 in the CONTROL treatment. Soil fertilized with urea increased the mean production of N2O to 0.015 µg N kg-1. Wastewater sludge increased the mean production to 0.11 µg N kg-1 and 0.58 µg N kg-1 in the L-SLUDGE and H-SLUDGE treatments respectively, and the vermicompost mean was 0.32 µg N kg-1 dry soil. N2O production in order of effect size of treatment was: wastewater sludge (H-SLUDGE) > vermicomposting > wastewater sludge (L-SLUDGE) > urea > unfertilized soil. Nitrifier nitrification and nitrifier denitrification were presumably the processes that contributed the most to the production of N2O under aerobic conditions.

#### **4.4. Fourth experiment under greenhouse conditions**

Juárez-Rodríguez et al. applied the sludge derived from anaerobically digested cow manure in the production of biogas (methane-air), to maize (*Zea mays* L.) cultivated in a nutrient-low, alkaline-saline soil with EC 9.4 dS m−1 and pH of 9.3. The results showed that the CO2 production increased 3.5-fold in the soil cultivated with maize and sludge, and increased 3.1 fold after the sludge was added to the soil. The production of CO2 from soil cultivated with maize showed a 1.6-fold increase compared with the uncultivated and unfertilized soil, 1.5 mg C kg−1 day−1. N2O production was -0.0004 µg N kg−1 soil day−1 in unfertilized soil, and in soil cultivated with maize was 0.3 µg N kg−1 day−1. Soil treated with sludge increased the produc‐ tion of N2O up to 4.6 µg N kg−1 soil day−1. Nevertheless, it was found that cultivated soil produced 2.4 µg N kg−1 soil day−1, reducing N2O production. It was also found that applying the anaerobically digested cow manure stimulated the growth of maize cultivated in an alkaline and saline soil, and the production of CO2 and N2O was increased.

#### **4.5. Fifth experiment under greenhouse conditions**

In this study, the main aim was to investigate how maize fertilized with wastewater at 120 kg N ha−1 affected crop growth, soil properties and the production of carbon dioxide (CO2), methane (CH4) and nitrous oxide (N2O) compared with plants fertilized with urea [16].

The soil was collected from the Mezquital Valley, located near Pachuca in the State of Hidalgo (Mexico). The irrigation water used was slightly alkaline with a pH of 8.4. The experiment was carried out under greenhouse conditions. Soil collected from three sub sites was placed into cylindrical pots. Five treatments were established in order to study the effect of wastewater and urea on the cultivation of maize (*Zea mays* L.). The treatments were: a) SMWW, maize plant plus wastewater; b) SMUREA, maize plant plus urea as fertilizer; c) SUREA, uncultivated soil and urea as fertilizer; d) SWW, uncultivated soil plus wastewater; and e) SCONTROL treatment, soil plus tap water. Soils from the SMWW and SWW treatments were irrigated with 1000 mL of wastewater every 7 days from the first day onwards, making a total of 13 times overall. This means that a total amount of mineral N equivalent to 120 kg N ha−1 was added to each maize plant, i.e. the recommended amount of N fertilizer for maize.

The concentration of NH4 <sup>+</sup> was larger in the soil treated with urea and wastewater than in the untreated soil, as the urea was hydrolysed and the wastewater contained high concentrations of NH4 + . The addition of wastewater to the soil doubled the production of CO2 and approxi‐ mately 0.2 g C was produced from the soil due to the decomposition of the wastewater after 70 days. In other words, 34% wastewater C was mineralized. However, urea may only occasionally stimulate CO2 production when a soil is N depleted. Plants take CO2 from the atmosphere, but mineralization of root exudates increases the production of CO2. The pro‐ duction of CO2 increased towards the end of the period of maize growth. This indicated that the phenological stage of the plants affected CO2 production. The growth of maize plants was similar under the SMWW (wastewater) and the SMUREA (urea) treatments, even when the release of nutrients was delayed by mineralisation from the organic matter in the wastewater. When wastewater was applied to the soil, the mean production rate of CO2 increased signifi‐ cantly at 2.4-fold, 1.7 µg C kg−1 h−1, compared with the SCONTROL treatment at 0.7 µg C kg−1 h−1 (Table 2). Meanwhile, cultivating maize increased CO2 production 3.2-fold, 5.6 µg C kg−1 h −1. The SWW, SMWW or SUREA treatments did not show a significant difference in the production of N2O compared with the SCONTROL (1.5×10−3 µg N kg−1 h−1). The addition of urea did not affect the CH4 oxidation rate (0.1×10−3 µg C kg−1 h−1), nor did the SMUREA treatment (cultivated soil fertilized with urea), but the addition of wastewater to the soil significantly increased CH4 production to 128.4×10−3 µg C kg−1 h−1. Soil irrigated with waste‐ water increased the global warming potential (GWP) up to 2.5-fold compared with the SUREA treatment (soil plus urea), whereas cultivated soil increased the GWP 1.4-fold. Crops irrigated with wastewater might limit the use of N fertilizer and water from aquifers. Nevertheless, the amount of fertilizer applied must be limited due to nitrate (NO<sup>3</sup> - ) leaching and the production of CO2, N2O and CH4 – that they could be produced in significant amounts –, and at the same time the salt content of the soil will accumulate, limiting the growth of the crop.

production increased 3.5-fold in the soil cultivated with maize and sludge, and increased 3.1 fold after the sludge was added to the soil. The production of CO2 from soil cultivated with maize showed a 1.6-fold increase compared with the uncultivated and unfertilized soil, 1.5 mg C kg−1 day−1. N2O production was -0.0004 µg N kg−1 soil day−1 in unfertilized soil, and in soil cultivated with maize was 0.3 µg N kg−1 day−1. Soil treated with sludge increased the produc‐ tion of N2O up to 4.6 µg N kg−1 soil day−1. Nevertheless, it was found that cultivated soil produced 2.4 µg N kg−1 soil day−1, reducing N2O production. It was also found that applying the anaerobically digested cow manure stimulated the growth of maize cultivated in an

In this study, the main aim was to investigate how maize fertilized with wastewater at 120 kg N ha−1 affected crop growth, soil properties and the production of carbon dioxide (CO2), methane (CH4) and nitrous oxide (N2O) compared with plants fertilized with urea [16].

The soil was collected from the Mezquital Valley, located near Pachuca in the State of Hidalgo (Mexico). The irrigation water used was slightly alkaline with a pH of 8.4. The experiment was carried out under greenhouse conditions. Soil collected from three sub sites was placed into cylindrical pots. Five treatments were established in order to study the effect of wastewater and urea on the cultivation of maize (*Zea mays* L.). The treatments were: a) SMWW, maize plant plus wastewater; b) SMUREA, maize plant plus urea as fertilizer; c) SUREA, uncultivated soil and urea as fertilizer; d) SWW, uncultivated soil plus wastewater; and e) SCONTROL treatment, soil plus tap water. Soils from the SMWW and SWW treatments were irrigated with 1000 mL of wastewater every 7 days from the first day onwards, making a total of 13 times overall. This means that a total amount of mineral N equivalent to 120 kg N ha−1 was added to

untreated soil, as the urea was hydrolysed and the wastewater contained high concentrations

mately 0.2 g C was produced from the soil due to the decomposition of the wastewater after 70 days. In other words, 34% wastewater C was mineralized. However, urea may only occasionally stimulate CO2 production when a soil is N depleted. Plants take CO2 from the atmosphere, but mineralization of root exudates increases the production of CO2. The pro‐ duction of CO2 increased towards the end of the period of maize growth. This indicated that the phenological stage of the plants affected CO2 production. The growth of maize plants was similar under the SMWW (wastewater) and the SMUREA (urea) treatments, even when the release of nutrients was delayed by mineralisation from the organic matter in the wastewater. When wastewater was applied to the soil, the mean production rate of CO2 increased signifi‐ cantly at 2.4-fold, 1.7 µg C kg−1 h−1, compared with the SCONTROL treatment at 0.7 µg C kg−1 h−1 (Table 2). Meanwhile, cultivating maize increased CO2 production 3.2-fold, 5.6 µg C kg−1 h −1. The SWW, SMWW or SUREA treatments did not show a significant difference in the production of N2O compared with the SCONTROL (1.5×10−3 µg N kg−1 h−1). The addition of urea did not affect the CH4 oxidation rate (0.1×10−3 µg C kg−1 h−1), nor did the SMUREA

. The addition of wastewater to the soil doubled the production of CO2 and approxi‐

<sup>+</sup> was larger in the soil treated with urea and wastewater than in the

alkaline and saline soil, and the production of CO2 and N2O was increased.

each maize plant, i.e. the recommended amount of N fertilizer for maize.

**4.5. Fifth experiment under greenhouse conditions**

The concentration of NH4

of NH4 +

220 Greenhouse Gases

Soils can be either a net sink or a net source of CH4, depending on several factors such as the moisture level, N level, and the nature of the ecosystem in question. Methane is used up by methanotrophic microorganisms, which are ubiquitous in several soils, and is produced by methanogenic microorganisms in the soil under anaerobic conditions. Agricultural systems are not normally large sources or sinks of CH4. Only under certain conditions are they sources of CH4 — after application of manure or other organic materials, or moderate to high levels of irrigation. Our results showed that soil irrigated with wastewater — with or without maize increased CH4 production significantly (SMWW and SWW treatments) particularly after irrigation, due to temporary anaerobic conditions.


<sup>a</sup> The global warming potential (GWP) of the gases produced was calculated considering CO2 production equivalent to 310 for N2O, 21 for CH4 and 1 for CO2 (IPCC, 2007) over a 90-day period, minus the C that was stored in the roots per kg soil.

b Values with the same letter show no significant difference between treatments (*P* < 0.05).

**Table 2.** Production of greenhouse gases, CO2, CH4 (µg C kg−1 soil h−1), and N2O (µg N kg−1 soil h−1) from five treatments: a) soil + plant + wastewater (SMWW), b) soil + plant + urea (SMUREA), c) soil + urea (SUREA), d) soil + wastewater (SWW), and e) soil + water (SCONTROL).

Fertilizing maize with urea or wastewater had a similar effect on plant growth, so wastewater might be useable as a crop fertilizer. The treatments with urea or wastewater had no effect on the pH of soil in this experiment due to the fact that the soil is a vertisol, characterized by a clay type 2:1, with a large capacity for the exchange of protons and consequently, a high buffering capacity. The addition of wastewater increased the production of both CO2 and CH4 compared with the soil treated with urea, but did not increase the production of N2O. The irrigation of crops with wastewater might in the long term be a far more environmentally friendly approach to that of using water from aquifers that take long periods of time to fill, as long as the amount of wastewater applied is restricted to the amount required by the cultivated crop due to possible substantial losses of mineral N through several process such as the production of CO2, CH4 and N2O, and the fact that soil salinization could increase rapidly.

#### **5. Conclusions**

The organic fertilizers or treatments (vermicompost, wastewater sludge, anaerobically digested cow manure, and wastewater) might increase the production of greenhouse gases, as do several abiotic and biotic factors involved in microbial activity within the soil. Sludge as a soil fertilizer offers the opportunity for the recycling of plants nutrients and the recovery of C as organic matter and its use in soil to improve agriculture. The high levels of microbial activity stimulated by the addition of this high C and N content material could increase both CO2 and N2O production. It should be taken in account that several factors are involved in the production of gases in the soil such as rhizosphere respiration, vegetation type, and soil microbial respiration, as well as abiotic factors such as soil moisture, soil temperature, substrate quantity and quality, and land use and management regimes.

Environmental and economic implications must be considered in order to make well-informed decisions on the management of soil treatments, i.e. how many, how often and what kind of organic fertilizer should be used in order to improve crop production and simultaneously limit soil deterioration and greenhouse gases production.

### **Acknowledgements**

We would like to thank the Instituto Politécnico Nacional and CONACyT for their financial support and the grant-aided support received.

#### **Author details**

Fernando López-Valdez 1\*, Fabián Fernández-Luqueño 2 , Carolina Pérez-Morales 1 and Mariana Miranda-Arámbula1

\*Address all correspondence to: flopez2072@yahoo.com

1 Agricultural Biotechnology Group, Research Centre for Applied Biotechnology, Instituto Politécnico Nacional, Tlaxcala, Mexico

2 Sustainability of Natural Resources and Energy Program, Cinvestav-Saltillo, Saltillo, C.P. Coahuila, Mexico

#### **References**

friendly approach to that of using water from aquifers that take long periods of time to fill, as long as the amount of wastewater applied is restricted to the amount required by the cultivated crop due to possible substantial losses of mineral N through several process such as the production of CO2, CH4 and N2O, and the fact that soil salinization could increase rapidly.

The organic fertilizers or treatments (vermicompost, wastewater sludge, anaerobically digested cow manure, and wastewater) might increase the production of greenhouse gases, as do several abiotic and biotic factors involved in microbial activity within the soil. Sludge as a soil fertilizer offers the opportunity for the recycling of plants nutrients and the recovery of C as organic matter and its use in soil to improve agriculture. The high levels of microbial activity stimulated by the addition of this high C and N content material could increase both CO2 and N2O production. It should be taken in account that several factors are involved in the production of gases in the soil such as rhizosphere respiration, vegetation type, and soil microbial respiration, as well as abiotic factors such as soil moisture, soil temperature, substrate

Environmental and economic implications must be considered in order to make well-informed decisions on the management of soil treatments, i.e. how many, how often and what kind of organic fertilizer should be used in order to improve crop production and simultaneously limit

We would like to thank the Instituto Politécnico Nacional and CONACyT for their financial

1 Agricultural Biotechnology Group, Research Centre for Applied Biotechnology, Instituto

2 Sustainability of Natural Resources and Energy Program, Cinvestav-Saltillo, Saltillo, C.P.

, Carolina Pérez-Morales 1

and

quantity and quality, and land use and management regimes.

soil deterioration and greenhouse gases production.

support and the grant-aided support received.

Fernando López-Valdez 1\*, Fabián Fernández-Luqueño 2

\*Address all correspondence to: flopez2072@yahoo.com

**5. Conclusions**

222 Greenhouse Gases

**Acknowledgements**

**Author details**

Coahuila, Mexico

Mariana Miranda-Arámbula1

Politécnico Nacional, Tlaxcala, Mexico

