This fortnightly newsletter was curated and edited by: J.W. Holloway and his Team

Synopsis

Climate change associated with emissions of greenhouse gases (GHG) resulting from human activities is noted as the defining human development issue currently (UNDP, 2008). Through the production of feed, growth of productive animals and the supporting herd, and the disposal of animal waste, livestock production contributes 8 to 18% (depending on the estimation method used) of global GHG emissions (O’Mara, 2011) and emissions from animal agriculture are expected to increase 17% by the year 2020 (USEPA, 2006). A significant reduction of GHG emissions can be achieved by intensifying animal production and improved production technologies in less developed countries and using novel mitigation practices in modern production systems (Smith et al., 2007; O’Mara, 2011).

Commentary

This limited series of the occasional e-letters are comprised of (4) four articles. They will appear fortnightly and are published during November and December, though they will be accessible through our social media pages.

Analysis

Manure Storage & Treatment

Greenhouse gas emissions from stored manure are primarily in the form of CH4 (due to anaerobic conditions), although N2O emissions can occur, and NH3 volatilization losses are often large. A direct way to avoid cumulative GHG emissions is to reduce the time manure is stored (Philippe et al., 2007; Costa et al., 2012). Increasing the time of manure storage increases the period during which CH4 (and potentially N2O) is emitted, and the emission rate, creating a compound increase (Philippe et al., 2007). Temperature is a critical factor regulating processes leading to NH3 (Sommer et al., 2006) and CH4 (Steed and Hashimoto, 1994) emissions from stored manure. Decreasing manure temperature to <10°C by removing the manure from the building and keeping it outside in cold climates can reduce CH4 emissions (Monteny et al., 2006). Storage treatments with proper aeration and moisture management have reduced CH4 generation from poultry manure (Li and Xin, 2010).

Ventilated belt removal of laying hen manure can reduce CH4 emissions compared to deep-pit storage (Fabbri et al., 2007). Separation of swine slurry into solid and liquid portions and then treating the solids through aerated composting reduced CH4 emissions by 99% and N2O emissions by 75% compared with untreated manure (Vanotti et al., 2008). However, due to the often negative relationship between NH3 and N2O emissions (Petersen and Sommer, 2011), this process is likely to increase NH3 emissions substantially and perhaps total N losses from manure. Amon et al. (2001) reported more significant NH3 losses from an actively turned composting pile of solid cattle manure than from an undisturbed anaerobically stored pile, with the opposite effect observed for N2O emissions. Prapaspongsa et al. (2010a,b) compared 14 swine manure management practices based on combinations of thermal pretreatment, anaerobic digestion, anaerobic co-digestion, liquid, and solid separation, drying, incineration, and thermal gasification concerning their energy, nutrient, and GHG balances.

The anaerobic digestion-based scenario with a natural crust during storage had the highest GWP reduction through high efficiencies in energy and nutrient recovery with restricted emissions of GHG and NO –. The incineration and thermal gasification-based scenarios and a scenario using only deep injection yielded the greatest reduction in respiratory inorganics and terrestrial eutrophication categories because they had the lowest NH3 emissions. Manure incineration combined with liquid and solid separation and drying of the solids was a promising management option yielding a high potential energy utilization rate and GHG reduction. Storage Covers. Several types of manure storage covers have been reported in the literature, including natural crusts on slurry manure stored with high solids content, straw, wood chips, oil layers, expanded clay pellets, wood, and semipermeable and sealed plastic covers. The effectiveness of the cover depends on many factors, including permeability, cover thickness, degradability, porosity, and management.

Semipermeable covers tend to increase N2O emission because they provide optimal aerobic conditions for nitrification at the cover surface and at the same time create a low oxygen environment just below the cover favorable for denitrification and the production of N2O (Hansen et al., 2009; Nielsen et al., 2010). Semipermeable covers are valuable for reducing NH3, CH4, and odor emissions, but they often increase N2O emissions (Sommer et al., 2000; Guarino et al., 2006; VanderZaag et al., 2008). Capturing the gases produced using impermeable membranes, such as oil layers and sealed plastic covers, can reduce NH3, N2O, and CH4 emissions. The results from Guarino et al. (2006) and VanderZaag et al. (2008) suggest that using a vegetable oil layer as a manure storage cover, although very effective, is not practical because of degradability, generation of foul odors, and difficulty in preventing the oil film from becoming mixed or “broken” over the manure surface.

Covering manure storages with impermeable covers is an effective mitigation practice if the CH4 captured under the cover is burned using a flare or an engine generator set to produce electricity; otherwise, the captured CH4 builds pressure inside the storage, creating an explosion hazard and/or rupture of the cover. Increased air pressure inside the storage structure reduces the fraction of compounds in the gas phase and increases that trapped in liquid manure. The increased gas trapped in the manure liquid is then released when the pressure in the manure storage is reduced to remove the manure. Retaining the CH4 produced is not beneficial if it escapes at a later stage; therefore, burning or combusting the collected CH4 to produce electricity or heat is the most desirable option. The effectiveness of impermeable covers depends on transforming the collected gases to less potent GHG such as NOx and CO2 (Nicolai and Pohl, 2004; Rotz and Hafner, 2011).

Anaerobic Digestion

Anaerobic digestion is the process of degradation of organic materials by archaea in the absence of oxygen, producing CH, CO, and other gases as by-products. This provides a promising practice for mitigating GHG emissions from collected manure. In addition, when correctly operated, anaerobic digesters are a source of renewable energy in the form of biogas, which is 60 to 80% CH4, depending on the substrate and operation conditions (Roos et al., 2004). Anaerobic digesters also provide opportunities to reduce pathogens and manure odor. Three practical temperature ranges are generally considered for anaerobic biogas systems: psychrophilic (15 to 25°C), mesophilic (30 to 38°C), and thermophilic (50 to 60°C). These temperature ranges facilitate the growth of specific microbes. Thermophilic systems are more sensitive to environmental changes, such as temperature fluctuations and chemical concentrations produced during the digestion process (Kim et al., 2002; Ahn and Forster, 2002; El-Mashad et al., 2003).

The number of functional microorganism species that thrive at this temperature is considerably less than those that survive at lower temperatures (Ziekus, 1977; Wolfe, 1979; Smith, 1980). Below 15°C, the production of biogas is greatly reduced and CO2 becomes the dominant product of anaerobic digestion; therefore, anaerobic digestion systems are not recommended for geographic locations with average temperatures below this threshold without supplemental heat and temperature control (Sommer et al., 2007). The effluent of the digester, commonly called digestate, contains most of the soluble plant nutrients found in the feedstock and the more resilient, difficult to degrade organic material. Digestate is commonly applied directly to crops whereas the sludge, formed by precipitated minerals and undigested OM, may be composted before field application. Digester designs vary widely in size, function, and operational parameters.

Smaller digesters (6 to 10 m3), designed to improve sanitary conditions in developing countries and to provide energy for single-family dwellings, were promoted in the 1970s and 1980s throughout Asia and Latin America (Bond and Templeton, 2011; Jiang et al., 2011b). These were designed to function with the waste originating from a few animals (2 to 5 swine, 5 to 10 cows, 100 chickens, or a combination of these) together with the family dwelling waste. According to Dhingra et al. (2011), these types of digesters reduced GHG emissions from 23 to 53% compared with households without biogas, depending on the condition of the digester, technical assistance, and operator ability. The effectiveness of these types of digesters for mitigating GHG depends mostly on the amount of CH4 leakage that occurs through digester walls and piping delivering the biogas to the family dwelling (Dhingra et al., 2011). These small digesters have been used by farmers in developing countries only when government subsidies and economic incentives have been available (Bond and Templeton, 2011).

Commercial farm digesters are typically designed to treat liquid manures. There are four basic commercial farm-level anaerobic digestion vessel designs (Roos et al., 2004). The most common and simple covered lagoon digestion systems are ambient (psychrophilic) temperature systems that require manure with a solid content of 3% or less and a storage cover to maintain anaerobic conditions. These systems typically create the largest type of digester with the longest hydraulic retention time. Plug-flow digesters and fixed domes use a vessel that receives manure at one end and discharges from the opposite end with no mixing or agitation. These systems may be heated to a mesophilic temperature and require slurry with a solids content of 11 to 13%. Small-scale digesters are often of this design. The more sophisticated complete mix digesters consist of an engineered digestion vessel designed to handle manure slurries with a solid content from 3 to 10%. A mixing system enhances bacterial contact with OM.

Supplemental heat is often added to these systems to operate at mesophilic temperatures, promoting bacterial growth and shorter hydraulic retention time. Fixed film digesters use a medium, such as rope, plastic mesh, or beads, placed in the vessel on which bacteria can grow. Dilute manures with a solids content of 3% or less are passed across (or through) the medium in these systems. Whereas other systems rely solely on suspended microbial growth, these also feature attached microbial growth. Widespread commercial farm digester adoption has not occurred because of variable economic return (Hill et al., 1985; Safley and Westerman, 1994; Braber, 1995) and the limited competitiveness of biogas with other fuels used for heat and power (Lantz et al., 2007). Industrial biogas digesters are used to produce renewable energy for towns and municipalities. These digesters, prevalent mostly in Europe, use biomass collected from several farms to feed the anaerobic digesters.

Co-digestion of agricultural biomass, industrial organic waste, and animal manures are common in industrial biogas plants. It allows better optimization of C to N ratio and CH4 production while reducing the impact of NH + on gas production (Ward et al., 2008). When CH4 is collected and used as an energy source, it can substitute for combusted fossil fuels reducing the emissions of GHG, NOx, hydrocarbons, and particulate matter (Börjesson and Berglund, 2006). These authors compared the emissions from the life cycle of raw materials used for the anaerobic digestion (6 different feedstocks, including swine manure) and the emission from systems that the anaerobic digestion process replaced. One of the serious concerns identified was uncontrolled losses of CH4 from biogas plants, including losses from stored digestate. Typical losses from systems storing digested manure were reported to range from 5 to 20% of the total biogas produced (Bjurling and Svärd, 1998; Sommer et al., 2001).

In a follow-up report, Börjesson and Berglund (2007) further explored the overall environmental impact when biogas systems replaced various energy-producing reference systems. The investigation was based on Swedish conditions using an LCA approach that considered both direct and indirect emissions. Greenhouse gas emissions per unit of heat were reduced by 10 to 25% when biogas-based heat replaced fossil fuel-based heat. Emissions from biogas systems contributed 60 to 75% and 25 to 40% of the life cycle emissions of CO2 and CH4 in the reference and fuel-based systems, respectively. During the anaerobic digestion process, N-containing compounds found in substrates, such as proteins, AA, and urea, are reduced to NH3 (Bernet et al., 2000). Ammonia remaining in the aqueous solution is then transferred to the soil when the digestate is land applied (Bernet et al., 2000; Hafner et al., 2006). Anaerobic digestion stabilizes the organic C in the feedstock (reducing the fraction of easily degradable C in manures) increases plant availability of N, and provides less energy to support the growth of N2O-forming microorganisms, reducing the potential of N2O emissions when applied to the soil.

Mineralization of organic N and VFA during anaerobic digestion increases manure pH and available N, resulting in increased NH3 volatilization (Petersen and Sommer, 2011). In general, reduction of manure OM content is expected to reduce N2O emissions from manure-amended soils (Petersen, 1999; Bertora et al., 2008) although Thomsen et al. (2010) reported higher N2O emissions when treated manure was applied in a wet spring season. These contradictory results led to Petersen and Sommer (2011) to conclude that there is no simple relationship between removing manure OM and the risk of N2O emission. To address this controversy, Thomsen et al. (2010) proposed linking the balance between N2O and N2 to soil water-filled pore space and oxygen supply. This relationship has been discussed in detail by Petersen and Sommer (2011); the authors concluded that the prediction of N2O emissions from manure-amended soil depends on manure composition and soil conditions.

Masse et al. (2011) noted high variability between N2O emissions, referring to 6 studies that found similar differences in emissions of the gas when comparing digested and nondigested manures. Data on the anaerobic digestion of poultry waste as a GHG mitigation practice are limited. Several studies show successful biogas production using poultry waste as a component of co-digestion (digesting poultry waste with other manures—beneficial due to the complimentary composition of the different manures); however, the impact on GHG mitigation was not reported. The ability to use anaerobic digestion to create, capture, and destroy CH4 derived from swine manure is well documented (Safley and Westerman, 1994; Masse et al., 2003a,b). Although it is possible to reduce CH4 emissions by over 60% from swine manure using anaerobic digestion, the amount of CH4 produced and collected does not directly translate into an equal amount of reduced CH4 emissions because the untreated manure would not yield the same amount of CH4 gas.

Most literature reviewed focused on research that compared digested manure with manures that received no treatment or a different treatment. In this manner, the biogas removed was not considered in the emission comparisons of nondigested versus digested manure by many authors who used the assumption that biogas produced during digestion is destroyed through controlled combustion. A number of the studies referenced here considered emissions from digested manure after it was land applied. Reductions of N2O emissions reported in these papers were as high as 70% compared with untreated manure applications. A commonly stated reason for this decrease was that digested manure contains less OM (degradable C), providing less energy for nitrite-forming microorganisms, which subsequently limits N2O production. Although most anaerobic digestion systems significantly reduce GHG emissions compared with traditional manure handling systems, incorrect operation, lack of maintenance, and CH4 leaks can make them a net contributor to GHG.

For this reason, system designs and components must ensure the containment of nearly all biogas. The potential for anaerobic digestion to mitigate N2O emissions after the digested manure is land applied is promising, but many parameters involved with field application contribute to conflicting reports. Anaerobic digestion systems require large initial capital investments during construction along with ongoing maintenance and supervision costs. Historically, the adoption of this technology occurs only when economic incentives are offered as price advantages for biogas (biofuels and renewable policy incentives), when the costs of construction and maintenance are subsidized, or when no competitive alternative energy source is available. Furthermore, instruction and technical assistance to users are necessary for implementing successful anaerobic digestion systems because the correct operation of anaerobic digesters is not trivial and 50% failure rates are typical (Bond and Templeton, 2011; Jiang et al., 2011a).

Acidification

An important factor affecting GHG emissions, in particular NH3, from stored manure is pH. According to Petersen and Sommer (2011), manure acidification is an effective mitigation option for NH3 emissions, but the effect on N2O is not well studied. The relationships between NH3 volatilization and factors such as air velocity and turbulence, manure temperature, and manure pH have been well documented (Ndegwa et al., 2011). Ndegwa et al. (2011) listed 15 studies in which cattle, swine, or poultry manure NH3 emissions were successfully mitigated (from 14 to 100% reduction in emissions) by lowering manure pH with sulfuric, hydrochloric, or phosphoric acids, calcium chloride, alum, or monocalcium phosphate monohydrate. These authors concluded that strong acids are more cost-effective at reducing manure pH than weaker acids or acidifying salts. However, strong acids are more hazardous and, therefore, acidifying salts and weaker acids may be more suitable for on-farm use.

Acidification of urine and, consequently, manure from cattle or monogastric farm animals has also been attempted using anionic salts, high dietary levels of fermentable carbohydrates, organic (benzoic) acids, or Ca and P salts (Ndegwa et al., 2008). A commercial system used on several farms in Denmark acidifies a portion of the manure with concentrated sulfuric acid to a pH of 5.5, removes a portion of the acidified manure equivalent to daily manure production, and returns the remaining manure to the storage facility (Sørensen and Eriksen, 2009). These authors concluded that NH3 volatilization from acidified cattle and swine manure was low after both soil incorporation and surface application. Petersen et al. (2012) studied the effect of acidification on CH4 (and NH3) emission from fresh and aged cattle manure during 3 mo of storage using the equipment described above.

Manure pH was adjusted to 5.5 with sulfuric acid, samples of manure were stored for 95 d, and NH3 and CH4 emissions were monitored. Manure pH increased gradually to 6.5 to 7 during storage. Acidification dramatically affected emissions, reducing CH4 by 67 to 87% (more pronounced with aged manure) and almost wholly eliminating NH3 emissions. The authors concluded that manure acidification might be a cost-effective GHG mitigation practice. Application of acidified manure is not expected to greatly impact crop production; the pH range of acidified manure is within the optimal range for corn and many cereal crops (5.5 to 6.5; Tisdale et al., 1993). Approximately 30% of soils worldwide and about 60% in Asia are acidic (<pH 5.5) and already require periodic lime applications to maintain optimal pH (von Uexküll and Mutert, 1995). Smaller quantities of acidified manure would be needed to provide crop N requirements because the reduction in NH3 emissions provides manure with a more significant plant-available N content.

However, long-term impacts of land application of acidified manures on soil pH have not been reported, and more frequent application of lime to maintain optimal pH in some soils may be required. Composting. Composting is an exothermic, aerobic process of microbial decomposition of OM that has several benefits related to manure handling, odor control, pathogen control, OM stabilization, additional farm income, etc. Composted manure solids (following manure separation into solids and liquid) are also used as bedding in some dairy production systems to reduce the cost of production and provide cow comfort, assuming udder health is not compromised (Husfeldt et al., 2012). Due to the nature of the composting process, N losses can be high and are influenced by a number of factors including temperature, C to N ratio, pH, moisture, and material consistency (Zeman et al., 2002). Compost can be a source of N2O emissions with both nitrification and denitrification processes occurring during composting.

Bacillus species are the main players in the degradation of OM and betaproteobacterial NH3–oxidizing bacteria involved in the nitrification process (Maeda et al., 2011). Depending on composting intensity, NH3 losses can be particularly high, reaching 50% of the total manure N (Peigné and Girardin, 2004).

Aeration of the composting heap reduces CH4 emissions (Thompson et al., 2004; Jiang et al., 2011a; Park et al., 2011) but can increase NH3 and N2O losses (Jiang et al., 2011a). Hao et al. (2004) reported up to 30% DM, 53% C, and 42% of initial N being lost during composting of straw-bedded manure. Methane losses accounted for 6% of the C losses. Nitrous oxide losses represented 1 to 6% of the total N losses. Addition of mature compost with nitrite-oxidizing bacteria to actively composting swine manure was shown to reduce N2O emission by 70% (Fukumoto and Inubushi, 2009).

These authors reported that up to 19% of the total manure N was lost as NH3 and N2O. Brown et al. (2008) reviewed the impact of composting of a range of feedstocks (including animal manure) on GHG emissions and pointed out that the primary benefit of composting is the reduction of CH4 emissions compared with manure stored under anaerobic conditions. These authors estimated, for example, that a facility that composts an equal mixture of manure, newsprint, and food waste could conserve the equivalent of 3.1 Mg CO2 per Mg of dry feedstock composted if feedstocks were diverted from anaerobic storage lagoons and landfills without gas collection mechanisms. According to Clemens et al. (2006), raw cattle manure can release from 160 (winter) to 3,600 (summer) g/m3 of CH4 and 38 to 57 g/m3 of N2O. For digested manure, the release rates are from 80 (winter) to 1,200 g/ m3 (summer) CH4 and 40 to 76 g/m3 N2O, respectively.

A recent study by Kariyapperuma et al. (2012) reported a 57% decrease in soil N2O emissions with composted vs. liquid swine manure. Remarkably, emissions during the same period of the following year were not different between composted and non composted manure; the authors attributed the lack of difference to a significant reduction in emissions in the second year due to frozen soil. In spite of significant GHG emissions from composting, the review by Brown et al. (2008) concluded that even in a worst-case scenario, these emissions are minimal in comparison to the benefits associated with the CH4 reduction credits from composting. The authors also stated that it is possible to significantly reduce emissions from compost piles by increasing the solids content of the feedstocks and the C to N ratio. Overall, Brown et al. (2008) concluded that composting could be an effective method for reducing GHG emissions from various waste materials, including animal manure. It must be noted, however, that NH3 losses during manure composting is significant.

Conclusion

In evaluating the position for red meat in the global environmental arena, an important consideration is that red meat makes a substantial contribution to food security providing essential nutrients and energy to human populations. Rumination allows ungulates to digest fibrous feeds that have little food value for humans, and thus, their contribution to food balances is significant. This contribution is of singular importance in marginal areas of the world, especially in the third world. The agro-ecological status and limited infrastructures offer few alternatives for food production. It is in these areas also that their value is apparent in that they can convert crop residues and fibrous by-products into high-quality edible products, and they contribute to soil fertility through their impact on plant nutrition and organic matter cycles (Gerber et al., 2015). Parallel to their contribution to the food chain, a full evaluation of their position in the environment necessitates evaluating their environmental sustainability issues.

These issues can be attributed to the computed low efficiency of beef cattle in translating natural resources (especially those that can be consumed by man) into edible products. For example, water use, land use, biomass appropriation, and greenhouse gas emissions are often computed to be higher per unit of edible product in red meat production systems than in any other livestock systems, even when adjusted statistically for product nutritional quality (Gerber et al., 2015). These computations usually fail to give beef cattle credit for converting highly lignified materials into high-quality human foods or consider that the feeds that cattle consume in competition to humans are utilized in production systems as “icing on the cake”; feeds that add value to beef from forage-fed production units (cow and calf units typically consume more than 70% of their nutrients from forages not consumable by man). Gerber et al. (2015) reviewed the literature concerning environmental challenges in beef production at the global level.

Beef production is faced with an array of additional sustainability challenges, including changing consumer perceptions, resilience to climate change, animal welfare, and inequities in access to land and water resources. Within the livestock sector, beef receives the most attention for its environmental impact. This results from the apparent aggregated contribution that beef production makes to global environmental issues such as climate change and land use (Gerber et al., 2015). In the last E-letter of this series, we will observe the plural solutions and hence applications that have been tested! We want to proceed further, more in-depth on this controversial subject. Therefore, please follow us on social media and join us on the (15^th) fifteenth of the month for Part (4!) four to learn more about the environmental impact of beef production. Thereafter, please join us on the (1^st) first and (15^th) fifteenth of each month for our fortnightly delivery of insightful, informative must-reads from some of the world’s scientific thinkers.

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