table of contents
The purpose of this paper is to provide a broad overview of …
Researchers have attempted to calculate PMN using laboratory methods, field studies, decay series, and computer models. Field studies by Cogger et al. (1999) found apparent biosolids N recovery of 28 to 40% for forage grasses and 11 to 44% for dryland winter wheat. Results indicated a greater availability of biosolids N in the second year of application than predicted from commonly used biosolids decay series (USEPA, 1983). Carryover N mineralized from biosolids applied to forage grasses can be significant; cumulative apparent N recovery increased by an average of 9% in the year following biosolids application (Sullivan et al., 1997). In a similar field study, Barbarick et al. (1996) reported net N mineralization from 25 to 57% for five to six applications of 6.7 Mg biosolids ha–1 on dryland winter wheat. More recent research from Barbarick and Ippolito (2000) found that 1 Mg of biosolids was equal to approximately 8 kg N fertilizer, whereas estimates calculated using the USEPA (1983) approach for the same material were only 6 to 7 kg N fertilizer equivalent per 1 Mg biosolids. They also estimated first-year mineralization rates that ranged from 25 to 32%.
Computer models and constant decay series have also been used to predict the amount of available N in municipal waste products. Gilmour and Skinner (1999) estimated PAN from six biosolids using decay-series modeling. This included average biosolids decomposition kinetic data from laboratory incubations, average field site weather data, and biosolids analytical data. Potentially available N released from biosolids during the field study was linearly related to the biosolids C to N ratio, organic N, or total N. Biosolids C to N ratio was the best predictor of biosolids PAN in the field, followed by organic N and total N. Slopes from the relationships of PAN to organic N and total N suggested that about 45 and 40% of the biosolids N was made plant available during the growing season, respectively.
Gilmour and Skinner (1999) used two approaches to calculate annual estimates of PAN from biosolids. For each approach, the computer model Decomposition employed analytical data for organic C, organic N, and inorganic N. One method used actual decomposition kinetics and weather data, and the other method used mean biosolids kinetics and mean weather data to estimate biosolids decomposition (Gilmour et al., 1996). Both methods produced similar estimates of PAN for the same biosolids source, but variability among biosolids was great. Mean estimates for PAN during the first year for Methods 1 and 2 were almost identical at 32.4 and 32.0 kg N Mg–1 biosolids, respectively. Overall, similar PAN predictions were obtained with decay series using the computer model with average or actual decomposition kinetics and weather data. This approach was verified with field studies located across a wide range of climatic conditions. Gilmour et al. (2003) reported that observed and predicted PAN were strongly correlated (r2 = 0.72), with the slope and intercept not significantly different than unity and zero, respectively (Fig. 1) . The application of this approach to other by-products has yet to be demonstrated.Accurate estimates of PMN are also needed for animal waste products. Land application of livestock manure can lead to surface and ground water contamination if rates exceed crop needs and reduce crop yields if rates limit N availability. Nitrogen mineralization in manure is influenced by a host of factors, both product-specific and environmental. Variability in N mineralization is due partly to soil type and field environmental conditions; however, the composition of the manure can significantly influence the rate and amount of organic N mineralization (Van Kessel et al., 2000). Net N mineralization in an incubation study ranged from –29.2 to 54.9% for several different dairy manure samples (Van Kessel and Reeves, 2002). While manure sample analysis found compositional differences in the samples, there was no strong correlation between various composition parameters and organic N mineralization (Van Kessel and Reeves, 2002). Van Kessel and Reeves (2002) identified no relationship between near-infrared spectral characteristics and N mineralization, but Qafoku et al. (2001) found a strong correlation (r2 = 0.82) between PMN and predicted N in poultry litter using near-infrared reflectance spectroscopy (Fig. 2) .
Haney et al. (2001) evaluated the relationships between the flush of CO2 during 1 d after rewetting of dried soil and potential soil N mineralization in the laboratory and forage N uptake in the field following dairy cattle manure application. The flush of CO2, or mineralized C, during 1 d was highly correlated with potential N mineralized in soil after a 24-d incubation and forage N uptake in the field. There was poor correlation between residual inorganic soil N and N mineralized during the same incubation. The C mineralized in 1 d after soil rewetting represents a rapid and reliable laboratory test to estimate potentially mineralizable N in manure-amended soils.
The potential for NH3 volatilization from many land-applied by-products, particularly poultry litter and anaerobically digested and alkaline biosolids, poses environmental and agronomic dilemmas. Concentrated animal feeding operations and wastewater treatment facilities can also be significant sources of NH3 emissions. Animal agriculture contributes approximately 50% of the total U.S. anthropogenic ammonia emissions (van Aardenne et al., 2001). Such N loss reduces the agronomic fertilizer value of the by-products and contributes to environmental problems such as acid deposition, eutrophication, reductions in biodiversity, and human and animal health concerns. In addition, NH3 exacerbates odor problems and reacts with other atmospheric constituents to produce haze. Conversely, NH3 volatilization could be viewed as advantageous by allowing greater residual application rates through reductions in the concentration of total N.
Quantification of NH3 emissions from various sources including CAFOs, wastewater treatment facilities, and land application of by-products is critical because of increased interest in the environmental effects of atmospheric NH3 and the lack of research on this topic. Unfortunately, the methodologies for measuring of NH3 emissions in the field limit our capabilities to predict N loss. Laboratory measurements are more useful for determining relative differences in NH3 volatilization potential than absolute field balances. Methods to accurately quantify NH3 losses in the field have not yet been developed.
Chemical amendments can reduce NH3 loss from poultry manure. Moore et al. (1995) tested several Al, Fe, and Ca amendments for reducing N loss from poultry litter via volatilization. Calcium hydroxide did not significantly affect NH3 loss in litter. Ammonia volatilization was reduced by 11 and 58% with high and low concentrations of FeSO4·7H2O, respectively, and by 24 to nearly 100% with alum + CaCO3 or alum alone. The higher alum rate resulted in a twofold increase of the N concentration in the litter, which increased the value of the material as an N fertilizer. Moore et al. (2000) also demonstrated that poultry health and production could be increased by alum reduction of NH3 emissions (Table 1).
Methods of application of by-product influence NH3 losses as well. Hansen et al. (2003) compared NH3 losses from six injection techniques with surface broadcast of liquid cattle manure and found subsurface placement reduced NH3 losses 20 to 75%.
Composting, a commonly employed organic waste processing method, can increase NH3 volatilization from manure depending primarily on the C to N ratio. Organic and inorganic amendments can also affect NH3 volatilization during the composting process. Kithome et al. (1999) evaluated the addition of adsorbents and chemical amendments to reduce NH3 emission during the composting of poultry manure. Adsorbents, such as zeolites, clay, and coir (fibrous material that makes up the thick mesocarp of the coconut fruit), are employed to bind NH3 and NH+4. Inorganic chemicals, such as CaCl2, CaSO4, MgCl2, MgSO4, and Al2(SO4)3, will inhibit the production of NH3. Ammonia loss from the unamended manure ranged from 14 to 29% of total N during the first 3 d and increased to 47 to 62% after 25 d. Manure amended with 38% zeolite and 33% coir reduced NH3 losses by 44 and 48%, respectively. The decreased volatilization resulted in greater extractable N in amended composts, thereby increasing the N fertilizer value. Alum also reduced NH3 volatilization, and increased the total N concentration in the resulting compost material.
Soil NO–3 accumulation and leaching to ground water are increased on land that routinely receives animal wastes. Nitrate from the Mississippi River basin has been implicated as one of the primary causes for the hypoxic zone in the Gulf of Mexico (USGS, 2000) (Fig. 3) . Efficient N management is vital for reducing N loss through both surface and subsurface transport. Recent interest in hypoxia and in N-based land application systems has generated interest in improving N fertilizer recommendation models to minimize nitrate losses from the landscape.
Andraski et al. (2000) evaluated four cropping–manure management system effects on soil water NO–3 concentrations and leaching below the root zone in corn. Early-season soil profile NO3–N contents varied among systems. Soil profile NO3–N contents increased by 80 to 210 kg N ha–1 from April to June in systems that had received past manure applications. The continuous corn system, with no manure history, had an increase of only 20 kg N ha–1. Management systems with a manure history had the greatest soil water NO3–N concentrations. Total NO3–N leaching amounts were greater for management systems with a history of manure application than for cropping systems with no manure history, and the economically optimum N rate was lower for sites with a history of manure applications (Table 2).
rating: 3.10 from 21 votes | updated on: 21 Jan 2007 | views: 12131 |