This fortnightly newsletter was curated and edited by: J.W. Holloway and his Team
Synopsis
Nitrous oxide emission occurs following land application as a byproduct of nitrification and denitrification processes in the soil. Still, these processes may also occur in compost, biofilter materials, and permeable storage covers. These microbial processes depend on temperature, moisture content,
availability of easily degradable organic C, and oxidation status of the environment, making N2O emissions and mitigation results highly variable. Managing the fate of ammoniacal N is essential to the success of N2O and CH4 mitigation because ammonia is a critical component for the cycling of N through manure, soil, crops, and animal feed. Manure application techniques such as subsurface injection reduce ammonia and CH4 emissions but can result in increased N2O emissions. The infusion works well when combined with anaerobic digestion and solids separation by improving infiltration.
Additives such as urease and nitrification inhibitors that inhibit microbial processes have mixed results but are generally effective in controlling N2O emission from intensive grazing systems. Matching plant nutrient requirements with manure fertilization, managing grazing intensity, and using cover crops are effective practices to increase plant N uptake and reduce N2O emissions. Due to system interactions, mitigation practices that reduce emissions in one stage of the manure management process may increase emissions elsewhere, so mitigation practices must be evaluated at the whole farm level.
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
Animal Housing
Structures used to house livestock do not directly affect the processes resulting in N2O and CH4 emissions; however, the type of structure used determines the manure management methods used to handle, store, process, and use the manure. Housing systems with solid floors that use hay or straw for bedding accumulate manure with higher DM, which is commonly stored in piles creating conditions conducive for nitrification and denitrification and thus greater N2O emission. Külling et al. (2001, 2003) compared liquid manure with stacked manure handling systems. Their results indicate that farmyard manure and deep litter manure handling systems tend to produce more significant N2O emissions than slurry-based systems. In these studies, quantitative differences in N2O emissions from the manure handling systems were difficult to determine because protein content in the diet and NH3 emission from manure also varied. Greater CH4 emissions were reported from farmyard manure followed by liquid slurry and deep litter manure. Amon et al. (2001) compared composted, anaerobically stacked, and slurry-based manure and found higher NH3 emissions in composted manure, with most of the losses occurring after manure was turned during aeration.
These authors found much greater N2O and CH4 emissions from anaerobically stacked manure with no significant difference between slurry-based and straw-based manure systems. Housing systems with slatted floors accumulate manure in liquid or slurry form with that manure stored for more extended periods increasing the production of CH4 and reducing the production of N2O. Hassouna et al. (2010) studied gaseous emissions from cattle housing in France. They found higher N2O emissions in buildings with straw-based bedding and solid manure handling systems when compared with liquid manure handling systems. Nitrous oxide emissions were detected in only 2 of the 14 liquid manure systems studied. The same study found smaller differences between CH4 emissions from buildings using straw-based solid manure or liquid manure systems. It attributed this result to the difficulty in discerning enteric CH4 emissions from manure emissions because the former produced most of the CH4 emitted from the buildings.
Hristov et al. (2012) investigated the effect of manure management on barn floor NH3, CH4, N2O, and CO2 emissions from 12 commercial dairy farms in Pennsylvania. Dairies participating in the study had flush (manure was flushed twice daily), two types of scrape (manure was scraped daily), and gravity-flow (manure was accumulated under the building and removed several times during the year) manure systems. Barn floor NH3 emissions were considerably lower for the flush manure systems (average of 167 mg/m2 per h) and highest for the gravity-flow system (426 mg/m2 per h). Methane emissions were also lowest for the flush (37 mg/m2 per h) and much higher for the gravity-flow system (1,216 mg/m2 per h). Carbon dioxide emissions were not different among manure systems (ranging from about 2,000 to 7,000 mg/m2 per h), and N2O emissions were negligible in all scenarios. This study showed that NH3, and particularly CH4, emissions from manure are much more significant from dairy barns where manure is stored for prolonged periods compared with barns where manure is removed daily.
In contrast to ruminants, housing plays a more critical role on GHG mitigation in nonruminant livestock production systems because most of the emission in these systems comes from the manure. Philippe et al. (2007) compared GHG emissions from fattening swine raised on a concrete slatted floor or straw-based deep litter. Swine fattened on deep litter released nearly 20% more GHG than those on slatted floors (6.2 and 13.1 g/swine per day for NH3, 0.54, and 1.11 g/pig per day for N2O, and 16.3 and 16.0 g/swine per day for CH4, respectively). The type of housing system also determines the feasibility of using anaerobic digestion or composting to treat the manure with its associated effects on GHG emissions. As described by the International Atomic Energy Agency (IAEA, 2008), the types of housing used in Asia facilitate different strategies for manure treatment depending on the capital resources of the producer. A popular alternative for smallholders is anaerobic digestion of animal and household waste, usually funded by government programs.
Medium to large producers have better access to capital investment, have specialized production determined by surrounding markets, and generally use raised slatted floors that allow the collection of manure for further treatment through solids separation and anaerobic digestion. Mechanically ventilated structures provide the opportunity to treat emitted GHG through filtration and scrubbing as gases are exhausted from facilities. An interesting mitigation technology for animal housing uses titanium dioxide (TiO2) paint on the interior walls. Industrial uses of TiO2 show that stimulation of its photocatalytic properties by UV light leads to oxidation of NH3 and NOx (e.g., Lee et al., 2002; Allen et al., 2005). Studies by Guarino et al. (2008) and Costa et al. (2012) in swine houses showed that GHG mitigation with TiO2 paint holds promise. Alkali and alkaline earth metal oxides, hydroxides, and carbonates or bicarbonates have been shown to have high CO2 absorption capacity and are being investigated for CO2 sorbent applications (Duan et al., 2012).
Dietary Effects on Manure Emissions
Diet manipulation to reduce nutrient excretion has been studied for many years, primarily related to N and P reduction. Still, its consideration as a mitigation practice to reduce CH4 and N2O emissions from manure are relatively new. Diet can profoundly affect N losses and particularly the route of N excretion (i.e., feces vs. urine) in most farm animals (Hristov et al., 2013b). Studies with 15Nlabeled urine or feces have demonstrated that urinary N is the primary source of NH + in cattle manure, contributing from 88 to 97% of the NH3 emitted within the first 10 d of manure storage (Fig. 3; Lee et al., 2011a). With urine being the primary source of volatile N emissions, manipulating the route of N excretion becomes an important N2O and NH3 mitigation tool. Urea is the main nitrogenous constituent of ruminant urine. In the urine of high-producing dairy cows, urea N represents 60 to 80% or more of total urinary N (Reynal and Broderick, 2005; Vander Pol et al., 2008) and proportionally decreases as dietary CP and intake decrease (Colmenero and Broderick, 2006).
On low-protein diets, urinary urea N can be as low as 46 to 53% of the total urinary N (Hristov et al., 2011a; Lee et al., 2012a) with urinary N close to or even below 19% of total excreta N (Lee et al., 2011b, 2012b). Therefore, reducing dietary CP concentration is perhaps the most effective method for mitigating NH3 emissions from stored manure (Hristov et al., 2011b). Emissions from land-applied manure are further reduced because low-CP diets produce manure with a slower mineralization rate of N (Powell and Broderick, 2011). Sauvant et al. (2011) showed that CH4 production per kilogram digested OM decreased linearly with increasing dietary CP; that is, decreasing dietary protein concentration likely results in increased concentration of fermentable carbohydrates in the diet, which in turn possible increases CH4 production. In support of this, Dijkstra et al. (2011) concluded, from a simulation study, that dietary-N mitigation options at the animal level aimed at reducing urinary N excretion may result in elevated enteric CH4 emission (per kg of fat and protein-corrected milk).
These potential relationships must be considered when manipulating dietary N to reduce manure NH3 and N2O emissions. Several studies have investigated the effect of dietary protein on N2O (and CH4) emissions from manure and manure-amended soil. Külling et al. (2001) reported decreased N2O emissions during simulated storage of manure from dairy cows fed low-protein diets. Still, the total GHG emissions were not affected by the dietary protein content (due to increased CH4 emissions from the low-protein manure). Velthof et al. (2005) concluded that decreasing the protein content of swine diets had the most considerable potential to simultaneously reduce NH3 and CH4 emissions during manure storage and N2O emission from soil. Data on the effect of dietary protein on manure N2O emissions, however, are not consistent and often no effect or even increased N2O emissions (from housing) have been reported when lowering dietary protein for swine (Clark et al., 2005; Philippe et al., 2006) and dairy cattle (Arriaga et al., 2010).
Manure CH4 and CO2 emissions per unit of land may increase immediately following soil application (or during storage) due to a more significant application rate for low vs. high-protein manure to meet the crops N requirements (Lee et al., 2012b). Feeding and management can significantly affect N excretions and volatilization losses from beef feedlots. Phase feeding is one example of an effective mitigation practice for these types of production systems. Reducing dietary protein concentration during the production cycle to better meet the requirements of the animal can significantly lower N excretions (Cole et al., 2005, 2006; Vasconcelos et al., 2007) and consequently losses from the pen surface. Erickson and Klopfenstein (2010) reported 12 to 21% lower N excretion and 15 to 33% lower N volatilization losses of phase-fed cattle. Manure management can also have a significant impact on N losses. Pen cleaning frequency, for example, decreased N volatilization losses by 19 to 44% and increased manure N by 26 to 41% (Erickson and Klopfenstein, 2010).
Reduced protein N in the animal diet produces manure with a slower N mineralization rate that releases less plant-available N (Powell and Broderick, 2011). Therefore, changes in manure application rate recommendations are needed to reflect N cycling from modified diets. At equal N application rates, whole-crop barley yield was not different between manures from dairy cows fed high(16.8) or low-crude protein (14.8%) diets (Lee et al., 2013). To minimize N2O production in all cases, manure application rates should be coordinated with the amount of mineral fertilizer applied, and consideration should be given to application timing and method to prevent N application over plant requirements. Low-protein diets must be formulated to meet or exceed the animal’s energy, metabolizable protein, and AA requirements if feed intake and animal performance are maintained (Lee et al., 2011b). Diets severely deficient in RDP will reduce total tract fiber digestibility in ruminants, which may negatively affect DMI and animal performance (Mertens, 1994; Lee et al., 2011b; Aschemann et al., 2012).
A meta-analysis by Nousiainen et al. (2009) and Huhtanen et al. (2009) showed that diet CP was the only dietary factor (of the factors studied in that analysis) that was positively related to NDF digestibility in dairy cows. Decreased ruminal degradability of fiber will increase excretion of fermentable OM in manure, which might increase manure CH4 emissions although the latter effect has not been consistently reported (Hindrichsen et al., 2005). On the other hand, these effects may be counteracted by reduced enteric CH4 production because fiber degradability in the rumen will decrease. Diets severely deficient in RDP will have a negative impact on microbial protein synthesis and animal productivity and, therefore must not be recommended as a mitigation practice. Feed intake depression with protein and AA-deficient diets has also been demonstrated with pigs and poultry (Henry, 1985; Picard et al., 1993) and must be avoided to maintain efficient animal production. Supplementation of low-protein diets with synthetic AA may alleviate undesirable effects on feed intake.
Growing swine fed a 14% CP diet containing supplemental lysine (0.73% Lys) had intake and growth performance similar to swine fed a 16% protein diet (0.77% Lys; Baker et al., 1975). Analogous results were reported by Yen and Veum (1982), who observed feed intake and ADG for growing swine fed a protein-deficient (13% CP) diet supplemented with Lys and Trp to be similar to swine fed a 16% CP diet. Similar to monogastric animals, supplementation of the diet with rumen-protected AA (Lys, Met and His) increased DMI and milk production in high-producing dairy cows (Lee et al., 2012a). Overall, feeding protein close to animal requirements, including varying protein concentration with the productive stage of the animals (phase feeding), is recommended as an effective manure NH3 and N2O emission mitigation practice. De Klein and Eckard (2008) concluded that the abatement of N2O should be considered part of an integrated approach to improve the efficiency of N cycling in animal production systems.
Particular attention should be given to improving animal N utilization, thus reducing urinary N output to the soil-plant system. According to these authors, current technologies could deliver up to a 50% reduction in N2O emissions from an animal housing system but only up to 15% from a grazing system. Dietary CP reduction can reduce both CH4 and N2O emissions from stored manure (Atakora et al., 2011a,b; Osada et al., 2011) and following land application (Velthof et al., 2005). All data do not support land application reductions because of the considerable variation in soil conditions. In intensively managed pastoral systems, supplementation of the pasture with low-N feeds such as corn or small grain silage, which generally reduces dietary N concentration, can reduce urinary N losses. Consequently, manure and soil NH3 and N2O emissions (by 8 to 36%; de Klein and Monaghan, 2011). In some systems, however, this reduction may be of smaller magnitude (Velthof et al., 2009), or total GHG emissions maybe even increased (Beukes et al., 2010), perhaps due to increased synthetic fertilizer use to grow the cereal silage.
Shifting N losses from urine to feces is expected to reduce N2O emissions from manure-amended soil due to the lower concentration of NH + in manure (this will depend on storage conditions if manure is stored before application). Feed additives, such as tannins, have redirected excreted N from urine to feces. Carulla et al. (2005), for example, reported a 9.3% reduction in urinary N as the proportion of total N losses, and Misselbrook et al. (2005) reported a decrease in manure NH3 emissions from cows fed stanniferous forage (although the CP content of the diets confounded the effect in the latter study). Grainger et al. (2009) observed a 45 to 59% reduction in urinary N excretion (as a percent of N intake) with condensed tannins but also a 22 to 30% drop in milk N secretion. Similarly, Aguerre et al. (2010) observed a linear decrease in urinary N excretion (vs. a linear increase in fecal N excretion) in high-producing dairy cows fed diets supplemented with 0 to 1.8% (DM basis) of a quebracho tannin extract.
Ammonia emission from slurry from cows receiving the tannin-supplemented diets was 8 to 49% lower than emissions from the control slurry. Tannins also reduced NH3 emission by 20% when directly applied to the barn floor and 27% after a tannin extract was applied to soil (Powell et al., 2011a,b). Studies directly investigating the effect of tannins on manure or soil N2O emissions are scarce. Hao et al. (2011) supplemented cattle diets with condensed tannins from Acacia mearnsii at 25 g/kg DM and followed GHG emissions from composted manure for up to 217 d. Nitrous oxide emissions that occurred during the first 56 d of composting were generally low (up to 0.1 kg N/t compost DM) and not affected by tannin supplementation. Methane (and CO2) emissions were also not affected by tannin supplementation. The authors speculated that the tannin application level was too low, tannins were complexed with protein on excretion in the feces, or that microbes in compost were capable of altering the biological activity of tannins.
Decreased N release rate from manure from animals fed stanniferous forages has been reported (Powell et al., 1994; Cadisch and Giller, 2001) although other reports indicated no effect of condensed tannins on the agronomic value of cattle manure (Hao et al., 2011) and decreased manure N availability may be a concern in agricultural systems relying exclusively on manure as a source of N for crop growth. Indeed, some reports had indicated a significant drop in yield when high tannin manure (equivalent to 2.2 t tannins/ha per yr) was added to sweet corn and radish plots (27 to 32% reduction for sweet corn and 42 to 46% for radish; Ingold et al., 2012). However, others reported no effect of condensed tannins on the N fertilizer value of feces from sheep receiving a diet supplemented with stanniferous legumes (Tiemann et al., 2009). More studies are needed to relate tannin application, through the diet or directly to manure, to GHG emissions from manure during storage or after land application.
Grazing Practices
Improving pasture quality in terms of forage digestibility is an efficient way of decreasing GHG emissions from the animal and the amount of manure produced. However, in pasture-based production systems, improving forage quality often means increasing N fertilizer application rates, which can have a negative impact on urinary N excretion and thus NH3 and N2O emissions. Nitrous oxide emissions can be exceptionally high in intensive pasture systems due to increased N concentration in urine due to the high CP content of pasture (22 to 28% CP, DM basis, in New Zealand, for example). There are many reports on the relationship between the placement and chemical composition of urine and soil nitrification and denitrification processes. Eckard et al. (2010) pointed out that the effective N application rate within a urine patch from a dairy cow on pasture is between 800 and 1,300 kg N/ha, and N is deposited at concentrations that are orders of magnitude greater than the utilization capacity of the soil-plant system.
These authors suggested that a more uniform distribution of urine throughout the paddock would reduce the effective N application rate, translating into a reduction in N2O emissions. These effects are often compounded by high fertilizer N application rates to stimulate grass growth, increasing urinary N concentration. De Klein et al. (2001) showed a 40 to 57% reduction in N2O emissions when grazing was restricted to 3 h/d in the humid late autumn in New Zealand. This reduction was attributed to diminished N input during conditions most conducive to N2O emissions in New Zealand. When de Klein et al. (2001) included N2O emissions resulting from applying the effluent collected during restricted grazing periods, N2O emissions were reduced by only 7 to 11%. Nevertheless, keeping the animals off the paddocks, in “stand-off” or “feed pads” for most of the day during the wet months of the year (autumn-winter), is an effective N2O mitigation practice in intensive grazing systems (de Klein, 2001; de Klein et al., 2002; Luo et al., 2008a).
Not allowing grazing during wet weather also increases pasture productivity due to reduced sward damage and soil compaction (de Klein, 2001; de Klein et al., 2006). One must keep in mind, however, that this practice results in much greater NH3 emissions (Luo et al., 2010) due to urine and feces being excreted and allowed to mix in the stand-off or feed pad area. According to Luo et al. (2010), reduction of N2O emissions from intensive grazing systems can be achieved by several strategies: 1) improving N use efficiency through reducing the amount of N excreted by grazing animals, 2) optimizing soil management and N inputs, 3) optimizing pasture renovation, 4) manipulating soil N cycling processes through soil additives, 5) selecting for plants and animals that maximize N utilization, and 6) altering grazing and feeding management.
Biofiltration
Biofiltration can treat ventilated air from animal buildings using biological scrubbers to control odor, to absorb NH3, and to convert NH3 into NO3. Preventing NH3 losses may indirectly reduce N2O emissions by reducing NH + deposition and consequent conversion to N2O (see earlier discussion). Ammonia removal efficiency in swine and poultry houses from acid scrubbers and biotrickling filters (based on biofilms that degrade the odorous compounds) averaged 96 and 70%, respectively (Melse and Ogink, 2005). Shah et al. (2011) investigated the effectiveness of a coupled biofilter–heat exchanger in reducing NH3 emissions (and recover heat) in a broiler house. The biofilter effectively treated very high inlet NH3 concentrations (>96 mg/kg) with removal efficiencies greater than 79% for empty bed residence times ranging from 4.3 to 29.1 s. The biofilter was also able to trap some sulfurous gases emitted from the broiler house.
Recent reports (Maia et al., 2012a,b) have shown that biofilters used to scrub NH3 from exhaust streams in animal houses generate N2O as a result of nitrification and denitrification processes in the biofiltration media. In their first study, Maia et al. (2012a) showed a high correlation between biofilter NH3 removal and N2O generation associated with converting NH3 to NO – and NO – in the biofilter. In their second study, Maia et al. (2012b) reported that moisture content between 48 and 52% in the biofilter media was an essential factor in obtaining significant NH3 reduction with reduced N2O production. Therefore, N2O production in biofilter scrubbers should be considered when implementing biofiltration systems for GHG and NH3 mitigation. A few studies have investigated CH4 mitigation bypassing contaminated air from a swine manure storage or from swine housing through a biofiltration system.
A Canadian Pork Council (2006) study reported reductions of 50 to 60%, and Girard et al. (2011) reported a maximum reduction of up to 40%. They described their biofilter as “packed with inorganic material,” but the packing material was not disclosed. Melse and van der Werf (2005) reported up to 85% CH4 removal from the exhaust stream of a covered swine liquid manure storage using a biofiltration system composed of a mixture of compost and perlite inoculated with CH4 oxidizing bacteria collected from activated sludge. The CH4 removal capacity of the biofilter system depended on the concentration of CH4 in the filtered stream. Therefore, the authors extrapolated that an equivalent system for animal housing with low CH4 concentration in the filtered stream would require extensive biofilter systems to achieve a 50% reduction or more, pointing to this as a limitation in the applicability of this technology.
In addition, they reported N2O production in the biofiltration system contributing 4 to 64% of the outlet stream GHG CO2e, which, as mentioned above, needs to be included in the design and promotion of biofiltration strategies for GHG mitigation. The high residence time necessary in these systems due to the low solubility and biodegradability of CH4 hinders effectiveness (Melse and Verdoes, 2005). Melse and Timmerman (2009) reported on the potential use of multipollutant scrubbers, combining acid scrubbers, biofilters, and water curtains to reduce not only NH3, odors, and GHG but also particulate matter from animal housing exhaust.
Conclusion
Worldwide, beef supply streams have been estimated to annually emit approximately 2.9 gigatons of CO2-eq, about 40% of all livestock emissions (Gerber et al., 2013). The greenhouse gas emissions per unit of product (emission intensity) is the highest when beef is produced on newly deforested land (Cederberg et al., 2011). Cattle are the primary ruminant species using about one-quarter of all emerged lands (Bouwman et al., 200 5). Assigning the total greenhouse emissions in these instances is an unfair burden to the red meat production industry. Some of the emission intensity is due to deforestation itself independent of the enterprise conducted on the deforested land. Our planet currently has over 1.3 billion cattle, approximately one for every five people (FAOSTAT, 2015). While the goals for cattle husbandry are quite varied delivering a wide array of products and functions, the vast majority is eventually harvested for beef (Gerber et al., 2015).
The debate over the environmental position of beef production is often characterized by a lack of recognition of this tremendous diversity in production and delivery systems, in the goods and services they deliver, and the environmental interactions and options for improvement that exist (Smith, 2015; and Herrero et al., 2013). In the following E-letter, we will examine other identified aspects that can be altered to reduce the GhG footprint! Therefore, please follow us on social media and join us on the (1st) first of the month for Part (3!) three to learn more about the environmental impact of beef production. Thereafter, please join us on the (1st) first and (15th) fifteenth of each month for our fortnightly delivery of insightful, informative must-reads from some of the world’s scientific thinkers. Selected by our editors is a collection of current topics with a profound ability for beneficial improvements, guidelines, and process practices.
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