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Effects of swine manure, chemical fertilizer, and drainage control on water quality Onwumere, George Chukwudi 1992

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E F F E C T S O F SWINE M A N U R E , C H E M I C A L F E R T I L I Z E R , A N D D R A I N A G E C O N T R O L O N W A T E R Q U A L I T Y by G E O R G E C H U K W U D I O N W U M E R E Diploma of Technology, B.C. Institute of Technology, Burnaby, 1982 B.Sc, University of Alberta, Edmonton, 1988 B.Sc, Special Certificate, University of Alberta, Edmonton, 1990 A THESIS S U B M I T T E D IN P A R T I A L F U L F I L M E N T O F T H E R E Q U I R E M E N T S F O R T H E D E G R E E O F M A S T E R O F S C I E N C E in T H E F A C U L T Y O F G R A D U A T E S T U D I E S B I O - R E S O U R C E E N G I N E E R I N G D E P A R T M E N T We accept this thesis as conforming to the required sta^darjd T H E U N I V E R S I T Y O F BRITISH C O L U M B I A July 1992 ©George Chukwudi Onwumere, 1992 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. 1 further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department The University of British Columbia Vancouver, Canada DE-6 (2/88) The global demand for agricultural cropland due to population growth and loss of prime cropland to soil erosion and urbanization has led to increases in chemical fertilizer usage in order to increase the yield per unit area. But this has not been without a price. Chemical fertilizers, apart from their high costs, are known to cause water pollution to both surface and subsurface water bodies. To mitigate this problem, research has been focused on the application of animal manures and composted material to land to improve the quality of some poorly drained soils. In this study, composted swine manure was manually mixed with a poorly drained soil sample collected from Boundary Bay Water Control project site to enhance its drainage. The soil plus compost mixture was introduced into clearcast acrylic tubes. Five of the tubes with soil alone control tube were set inside the Laboratory and the other five tubes with a control were situated outside to simulate field conditions. The tubes were fertilized with different concentrations of liquid swine manure and chemical fertilizer to supply the necessary nutrients required by plant. Enough water was applied to saturate the sample based on the soil water storage capacity from a previous experiment. Leachate samples were collected and analyzed for some selected water quality criteria (pH, electrical conductivity (EC), ammonia-N, nitrate + nitrite-N, and total Kjeldahl nitrogen). The p H , E C , ammonia-N, and nitrate + nitrite-N did not show any significant increase with the increased rate of liquid swine manure or chemical fertilizer application. In fact, the soil plus compost alone mixture may contain enough nutrients that extra fertilization was not necessary. Hence, the composted swine manure can be used as an extra source of nutrient and to enhance the soil hydraulic properties. Water pollution from these applications was minimal based on the experienced conditions. Abstract i i Table of Contents i i i List of Tables vi List of Figures vii Acknowledgement ix 1. Introduction 1 1.1 Population Growth and Chemical Fertilizer Usage 1 1.2 Water Quantity and Quality 3 1.3 Animal Manure Application Concerns and Remedies 7 1.4 Purpose of this study 8 2. Literature Review 11 2.1 Different Forms of Nitrogen 11 2.2 The Nitrogen Cycle 13 2.2.1 Nitrogen Inputs 13 2.2.2 Soil Organic-Inorganic Transformations 18 2.2.3 Nitrogen Removal Mechanisms 23 2.3 Soil Nitrogen Loading Calculation from Liquid Manure 29 2.4 Nitrate Leaching Models 33 2.5 Nitrate Seasonal Leaching Patterns 36 2.6 Fertilization 38 2.7 Water Table Manipulation Using Drainage Control 42 2.8 Summary 42 3. Materials and Methods 44 3.1 Experimental Overview 44 3.2 Site and Soil Description 45 3.3 Experimental Apparatus 47 3.4 Composted Swine Manure Used 47 3.5 Liquid Swine Manure and Chemical Fertilizer Used 51 3.6 Run Procedures 52 3.6.1 Run 1 53 3.6.2 Run 2 54 3.7 Nutrient Application Rates 55 3.8 Chemical Analysis 56 4. Results and Discussion 58 4.1 p H Measurements 58 4.2 Electrical Conductivity (EC) 59 4.3 Ammonia-N Analysis 70 4.4 Nitrate + Nitrite-N Analysis 84 4.4.1 Factors that Affected NO^ + NOj^-N Concentrations 94 4.5 Total Kjeldahl Nitrogen (TKN) 105 4.6 Drainage control Effects on NO^ + N0(2-N 105 5. Conclusion 107 6. Bibliography I l l 7. Appendix A Calculations 117 1- 1 The Global Distribution of Water 3 2- 1 Global Nitrogen Input and Output Mechanisms 43 3- 1 Soil Physical and Chemical Characteristics 46 3-2 Some of the Characteristics of the Liquid Swine Manure and Chemical Fertilizer 52 3- 3 Moisture Content of the Soil and Compost Before Mixing 54 4- 1 Inside p H Measurements for Run #1 60 4-2 Outside p H Measurements for Run #1 61 4-3 Inside p H Measurements for Run #2 62 4-4 Outside p H Measurements for Run #2 63 4-5 Inside E C Measurements for Run #1 66 4-6 Outside E C Measurements for Run #1 67 4-7 Inside E C Measurements for Run #2 68 4-8 Outside E C Measurements for Run #2 69 4-9 Inside Ammonia-N Concentrations for Run #1 72 4-10 Outside Ammonia-N Concentrations for Run #1 73 4-11 Inside Ammonia-N Concentrations for Run #2 74 4-12 Outside Ammonia-N Concentrations for Run #2 75 4-13 Inside NO^ + N O 2 - N Concentrations for Run #1 89 4-14 Outside NO^ + NOj^-N Concentrations for Run #1 90 4-15 Inside NO3 + N O 2 - N Concentrations for Run #2 91 4-16 Outside NO^ + NC^-N Concentrations for Run #2 92 1-1 World Population Trend from 1950 - 2000 4 1-2 World Fertilizer Production From 1946 - 1986 5 1-3 Swine Manure Production in Canada in 1976 and 1980 10 3-1 Schematic of the Experimental Setup 48 3-2 Dimensional Analysis of the Setup for Tubes #1-5 49 3- 3 Particle Size Distribution for Compost 8 B in #2 50 4- 1 Ammonia-N Concentrations for Tube #1 Run #1 78 4-1A Ammonia-N Concentrations for Tube #1 Run #2 78 4-2 Ammonia-N Concentrations for Tube #2 Run #1 79 4-2A Ammonia-N Concentrations for Tube #2 Run #2 79 4-3 Ammonia-N Concentrations for Tube #2 Run #1 80 4-3A Ammonia-N Concentrations for Tube #3 Run #2 80 4-4 Ammonia-N Concentrations for Tube #4 Run #1 81 4-4A Ammonia-N Concentrations for Tube #4 Run #2 81 4-5 Ammonia-N Concentrations for Tube #5 Run #1 82 4-5A Ammonia-N Concentrations for Tube #5 Run #2 82 4-6 Ammonia-N Concentrations for Tube 36 Run #1 83 4-6A Ammonia-N Concentrations for Tube #6 Run #2 83 4-7 Inside N I % - N Concentrations for Tubes #1-6 Run #1 85 4-7A Inside N H j - N Concentrations for Tubes #1-6 Run #2 85 4-8 Outside N I % - N Concentrations for Tubes #1-6 Run #1 86 4-8A Outside NH3 - N Concentrations for Tubes #1-6 Run #2 86 4-9 NO^ + N q 2 - N Concentrations for Tube #1 Run #1 95 4-9A NO3 + NO2 -N Concentrations for Tube #1 Run #2 95 4-10 NO^ + N O 2 - N Concentrations for Tube #2 Run #1 96 4-lOA NO^ + N O 2 - N Concentrations for Tube #2 Run #2 96 4-11 NO^ + N O 2 - N Concentrations for Tube #3 Run #1 97 4-1 l A N O 3 + N 0 2 - N Concentrations for Tube #3 Run #2 97 4-12 N O 3 + N O 2 - N Concentrations for Tube #4 Run #1 98 4-12A NO^ + N O ^ - N Concentrations for Tube #4 Run #2 98 4-13 NO^ + N O 2 - N Concentrations for Tube #5 Run #1 99 4-13A N O j + N O 2 - N Concentrations for Tube #5 Run #2 99 4-14 NO3 + NO2 -N Concentrations for Tube #6 Run #1 100 4-14A NO3 + NO2 -N Concentrations for Tube #6 Run #2 100 4-15 Inside NO3 + ^C^-N Concentrations for Tubes #1-6 Run #1 . . 101 4-15A Inside NO^ + NO2 -N Concentrations for Tubes #1-6 Run #2 . . 101 4-16 Outside N G ^ + NO2 -N Concentrations for Tubes #1-6 Run #1 . 102 4-16A Outside N O j + NO2 -N Concentrations for Tubes #1-6 Run #2 . 102 Acknowledgement I would like to thank Dr . S.T. Chieng (my Supervisor) for his assistance in the preparation of this report. His ideas, support, suggestions, and experience in the field of soil and water engineering ensured the success of this report. My thanks to Dr. K . V . Lo and Dr. S.O. Russell for reviewing the initial manuscript, providing me with the necessary advice, and sitting on my Committee. To J . Pehike and N . Jackson (Technicians), A . Chen, Y . Gao and Dr. P. Liao (Laboratory), I express my appreciation for their valuable help in the experimental setup and laboratory analysis. Also my sincere gratitude to Dr. Steve Smith (S.F.U.) for his constructive criticism and inspiration in preparing the final document. I would like to thank Ms. Kathy Lasuik for not only typing the document but also for her relentless support. Last, but not the least, my thanks to my family for being there and to friends for their understanding. I N T R O D U C T I O N Water quality problems were associated in the past with only the developing countries of the world. But increasingly, water quality degradation is also a menace to the developed countries. Most countries of the world have in fact, been negligent in the management of their water. To mitigate this problem, recent attention has been given not just to the symptoms of poor water quality but also to actually addressing the real causes of water quality deterioration. Water as a subsystem can undergo natural self-purification, which is a process of waste degradation by microbes in the water body. But the tendency for water bodies to undergo natural self-purification has declined very remarkably due to high population density in catchment areas, growth in industrialization, and acceleration in global food requirement. As a result, there is a net accumulation of waste substances in the oceans, lakes and rivers. These waste substances, in turn, not only pose a threat to aquatic life but also affect human water consumption patterns and the quantity of water available for agricultural food production. In effect, reducing population growth within the catchment area and improving chemical fertilizer efficiency would help enormously in ameliorating water quality standards. 1.1 Population Growth and Chemical Fertilizer Usage Population growth creates an accelerated demand for agricultural food production. As the global population increment passed the 80 million per year mark in 1980, the addition is more so in the developing countries than the developed countries of the world. Also, with the global population projected to be over six billion people in the year 2000, there will be tremendous pressure on agricultural cropland to sustain the populace. See Figure 1-1 (Brown et al., 1988). In most of the developing countries, due to rapid population growth, excessive cultivation and poor land management, there has been a decrease in per capita food production per unit area. With the availability of chemical fertilizers, however, some of the countries have been able to turn their declining food production around. O n the other hand, the increasing costs of chemical fertilizers have not helped poor developing countries. As a result, such countries depend on food importation from the developed countries. In developed countries such as Canada the population growth has either reached the zero mark or is in a declining phase. Most of the food products are for export market. A n d since there is a demand for the increased yield, chemical fertilizer application has been used extensively. Figure 1-2 depicts global chemical fertilizer production between 1946 and 1986. But this has not been without a price tag. Chemical fertilizers are known to pollute water bodies apart from their high cost. As a result, they not only endanger human and aquatic lives but also limit the quantity of water available for other purposes ( F A O , 1987). In the Lower Fraser Valley region of the province of British Columbia, for example, there has been a net increase in the population growth for the period 1980 - 1990 followed by a corresponding increase in food demand. This increase in population is due primarily to two things namely interprovincial migration and increase in foreign immigrants to the province. As a result, the province has to produce more food for both the local population and the export market. Accompanying this increase in population and foreign demand is a corresponding increase in the chemical fertilizer application to land. There is no statistical data to show the exact amount of chemical fertilizer application to croplands in this region. However, the decHning water quaUty standard is telling the farming community that it is time to not only curtail chemical fertilizer usage but also to look for an acceptable replacement. 12 Water Quantity and Quality Globally, there is enough water available for the world's five billion plus population, however, the highly uneven distribution of water and the increase in water pollution hinders water availability. Globally, approximately 97.41 percent of the earth's water is in the ocean with the remaining 2.59 percent being on land. Only about 0.01 percent of the 2.59 percent exists as the earth's total supply of fresh water (see Table 1.1). Of the 2.59 percent, the bulk of it is locked up as polar ice caps and glaciers with the remaining 0.60 percent being subsurface water (La Riviere, 1989). Table 1.1 Global Distribution of Water Form Percentage Oceans 97.41 Glaciers/Ice Caps 1.98 Subsurface 0.60 Lakes/Soil Moisture/Atmosphere/Rivers 0.01 Source: L a Riviere, 1989. On a continental basis, Asia, Oceania, and Africa are already having problems with water scarcity due to rapid population growth, increase in irrigation demand and drought. Europe and North America are better off, with South America being the most richly endowed (Falkenmark, 1989). 7 P O P U L A T I O N B I L L I O N 1950 1960 1970 1980 1985 YEAR 1990 1995 2 0 0 0 •I P O P U L A T I O N F I G U R E 1-1. W O R L D P O P U L A T I O N F R O M 1 9 5 0 - 1 9 9 0 W I T H P R O J E C T I O N S UP T O Y E A R 2 0 0 0 ( D A T A F R O M B R O W N E T A L „ 1 9 8 8 : FERTILIZER PRODUCTIONS(MILLION TONNES) • 1 WORLD PRODUCTIONS FIGURE 1-2, WORLD NITROGEN FERTILIZER PRODUCTION BETWEEN 1946 AND 1986 (DATA FROM FAD,1987) Canada is well endowed with water resources with the exception of the prairie provinces. The average annual precipitation for Canada is about 600 mm. The annual precipitation ranges from about 100 mm in the Arctic to well over 3,600 mm along the Pacific region. Most of this precipitation falls as rain with only about one-third falling as snow (Bertrand et al., 1985). British Columbia (B.C.) is probably one of the most water-rich provinces in Canada. British Columbia has an annual average precipitation that ranges from 1,000 mm to well over 2,000 mm, depending on the location. B .C . is also blessed with both surface and subsurface water bodies (Russell et al., 1979). As a result, the water resources should be able to withstand both the rapid population growth and the increasing food production requirement. It is not however, the water quantity per se that is the problem in B.C. , but the quality of the water that is available. Water bodies are susceptible to some form of pollution either from natural sources or from human activities or both. But increasingly, that susceptibility to pollution has been exacerbated by human activities. Municipal sewage, industrial wastes and nonpoint source pollution such as agriculture are all sources of water pollution caused by human activities. Municipal sewage and industrial wastewaters are both major sources of pollution and are both major concerns. But with the growth of environmental awareness, actions are being taken to curb these types of pollution both locally and internationally. Nonpoint source pollution from land-based activities such as agriculture is on the rise and its control is proving to be more difficult. Agriculture is a major nonpoint pollution contributor. Soil erosion is a serious problem because eroded material contains chemical fertilizers and pesticides which create heavy oxygen demand in the water bodies. These chemicals can also be very harmful to humans at elevated concentrations. This is a problem not only for water quality management, but also for the agricultural industry because soil loss threatens the land's productivity and forces farmers to use more fertilizers. The more fertilizers the farmers use, the higher the concentration of these organic chemical pollutants that reach the receiving water bodies. Consequently, the quality of the water is further degraded. In the Lower Fraser Valley there has been a tremendous increase in agricultural activities and chemical fertilizer application; however, the actual water quality degradation from nonpoint sources remains an educated guess. 1.3 Animal Manure Application Concerns and Remedies Even though land application as a method for "disposing" of animal manures and returning nutrients back to the soil has been the subject of a number of research projects, the environmental consequences are still much debated. Protecting the quality of the water bodies has emerged as a major issue. The principal concern is nitrate which is associated with both chemical fertilizers and other organic materials such as food processing wastes, animal manures, and sewage sludges. Nitrate accumulation in the groundwater has been suggested to be the major causes of several human health problems such as birth defects, cancer, nervous system disorder, and methaemoglobinaemia. Of all the illnesses, only the methaemoglobinaemia effect has been verified and well documented (Hubbard et a l , 1991; Moore, 1989 and Health and Welfare Canada, 1980). Methaemoglobinaemia, which has caused a number of infant deaths, is caused by alteration of the nitrate to nitrite in the digestive tract. The absorption of nitrite into the bloodstream and its interaction with the haemoglobin in the blood leads to methaemoglobin production. Methaemoglobin inhibits oxygen carriage to body cells, thereby leading to oxygen deficiency (Aldrich, 1980). Since chemical fertilizers are known to release high nitrate concentrations, using slow nitrate releasing organic materials such as swine manure could be encouraged. Organic materials such as animal manures and food processing wastes are valuable in agriculture for their properties as soil conditioners and fertilizers. They also contain nitrate and have the potential for contaminating the water bodies, but the nitrate concentrations are released slower in most cases. As a result, nitrate contamination from swine manure application may be considerably lower than contamination from chemical fertilizers (Sutton et al., 1978). Bacterial contamination may be of some concern with swine manure waste application. But if the swine manure is subjected to some kind of treatment before application, bacterial contamination should not pose any hazard to human health or water quality. Indeed, proper treatment of swine wastes before land application could minimize any form of bacterial contamination, and contribute enough nitrate to sustain crop yield without contaminating our water systems. 1.4 Purpose Of This Study The maintenance of basic water quality standard has come to be a very important focus in water resource management. In order to achieve this basic water quality standard, steps must be taken to improve those land-use activities that degrade water quality. The use of animal manure application is encouraged not only because of its use as a soil conditioner and fertilizer, but also as a waste reduction measure. In the past twenty years, environmental awareness has led to an increase in waste management which encourages the four R's - recycle, re-use, removal, and reduction of waste. Animal wastes application to land will be doing two things - waste removal and water quality maintenance or improvement. A t the present time, waste removal is an important issue in Canada since the total swine production has gone from 12 x IC^ tonnes (wet weight) to over 16.9 x IC^ tonnes from 1976 to 1980. Figure 1-3 shows total manure production on a province by province basis for 1976 and 1980. Most of the provinces including British Columbia show a substantial increase in swine manure production between 1976 to 1980 with the increase extending into the 1990's. As a result, removing this manure or applying it to land has become the subject of recent researches (Agriculture Canada, 1981). The main purpose of this thesis is to evaluate the potential for land application of swine manure at a higher application rate as opposed to chemical fertilizers in maintaining water quality standard. Other objectives are: 1. To investigate the effects of soil-compost mixture alone on water quality. 2. To generate data for p H , E C , ammonia-N, and nitrate + nitrite-N under different application rates. 3. To make inferences on practical application based on the observed data. 4. To compare the inside and outside Laboratory data for any correlation. MANURE PRODUCTION (MILLION TONNES) NEWFD P.E.I N.SCO N.BRU MARI QUEBEC ONT MAN SASK ALTA B.C. PRO/INCES YEAR •11976 Hiigso FIGURE 1-3. SWINE MANURE PRODUCTIONS IN CANADA IN 1976 & 1990 (DATA FROM AGRIC. CANADA PROCEEDINGS DEC.9-10,1981 Chapter 2 LITERATURE REVIEW There has been quite an extensive research done on the impact of different forms of nitrogen on water quality. The inorganic nitrogen contained or released from wastewater and sludges may well be the most important factor limiting the land application of wastes. High concentrations of nitrogen, while stimulating plant growth, can lead to surface and subsurface water contaminations. High concentrations of nitrogen in conjunction with phosphorous in the surface water body can lead to accelerated eutrophication or excessive growth of algae. Similarly, nitrate accumulation in drinking water supplies has been shown to be a health hazard to humans, especially infants. As a result, there is an existing guideline to maintain the subsurface and surface drinking water nitrogen level below 10 m g / L NC^-N (Rosswall and Paustian, 1984; Lucas and Reeves, 1980; and Zwerman, et al., 1972). In order to achieve this, nitrogen loading rates to the soil must be carefully monitored and scrutinized. A l l the forms of nitrogen available to the soil and their transformations must be known and thoroughly managed to avoid nitrate overload. 2.1 Different Forms of Nitrogen There are six different forms of nitrogen that are of importance in land application of waste, namely organic nitrogen (R-NH2), ammonical nitrogen (NH^"'' and N I ^ ) , nitrite (NO2'), nitrate (NOj'), nitrous oxide (f»^0), and molecular nitrogen (N^) (Loehr et al., 1979). Organic nitrogen, a constituent of humus, is not available for plant uptake within the soil profile. But this can be changed by organic nitrogen undergoing a slow transformation process. There are two forms of ammonical nitrogen: the ammonium ion (NI^"^ and anmionia gas (NI^)) . The positively charged ammonium ion can be held in the soil, especially at p H less than 8.5. A t high p H values ammonia gas is dominant (Lx)ehr et al., 1979). Nitrite (NO^') is very unstable, but may accumulate under certain conditions. This mobile anion is very toxic to plants. Nitrite is generally formed during an intermediate stage of ammonium conversion to nitrate (Moore, 1989). 2 NH4+ + 3 O2 ^ 2 NO2" + 2 B2O + 4 2 NO2" + ^2 2 NO^" (rapid oxidation to nitrate) Nitrate (NO^") is another highly mobile anion ion and is the end product of the nitrification reaction as can be seen from the above two equations. Nitrate is the nitrogen form that is of most concern because of not only the ease with which it can be leached from the soil but also because of its toxicity to humans, particularly infants (Miranowski, 1983 and Moore, 1989). Nitrous oxide (N2O) which is formed during denitrification reaction is rarely used by plants. This gas is usually present in the atmosphere and has a residence time of 100 - 200 years. It has a very low solubihty in water and hence poses no immediate threat to humans (Loehr et al., 1970, and Jenkinson, 1990). Finally, molecular nitrogen (N2), another end product of denitrification reaction, is present in the atmosphere at a very high percentage of approximately 79%. This gas also does not pose a threat to humans. Of all these forms of nitrogen, the nitrogen cycle in the soil used for agricultural purposes involves three major forms of nitrogen, namely organic nitrogen (R-NH^), ammonical nitrogen (NU^^ ), and nitrate nitrogen (NO^-N). Understanding the entry, the transformations, and the removal of these nitrogen forms would be crucial in maintaining an adequate water quality. 22 The Nitrogen Cycle The nitrogen cycle has been the subject of extensive research in the last 100 years. Though there are numerous literatures on this subject, researchers are still divided on how much nitrogen enters, transforms within, or leaves the system. Since this information is vital to estimating potential nitrate pollution, much of the discussion will be focused on nitrogen inputs, organic - inorganic transformations, and nitrogen removals. 2.2.1 Nitrogen Inputs Nitrogen input to the land comes from waste/fertilizer application, precipitation, and through fixation of atmospheric molecular nitrogen by microorganisms. How much nitrogen is introduced by each of these three input systems is still a subject of much debate, but Jenkinson (1990) estimated the total nitrogen input to be 290 million tonnes per year. His total estimate includes 74 million tonnes per year from fertilizer/waste application to land. Waste/Fertilizer Applications. Animal waste and fertilizer contain different amounts of nitrogen in the forms of organic-N, ammonium-N, and nitrate-N. How much is present of each depends on soil texture and the climate of the location. The total amount of nitrogen present depends, however, on the loading rate and the type of soil. In the United States, domestic livestock contribute about 6 million metric tons of nitrogen per year (Aldrich, 1980). In Canada, the total nitrogen contribution is not known but there are government recommended application rates which vary from province to province. Much of the research on application rates and their effects on water quality has been centered on these government recommended application rates. Research to date has indicated that total nitrogen content increases with rate of waste application. The work of Sutton et al. (1986), Mesk et al. (1974), Mugwira, (1979), Smith et al. (1980), and Copper et al. (1984), supports the notion that surface soil total nitrogen content is increased by the application of either cattle or dairy manure. Most nitrogen contributions are in the form nitrate-N. Sutton et al. (1978) found that nitrate-N concentrations increased with increasing liquid swine waste application rates. A similar result is obtained by Chang et al. (1991) after eleven annual appUcations of cattle feedlot manure. In the same study total nitrogen content is significantly affected by waste application in both irrigated and non-irrigated land (Chang et al., 1991). But there seems to be a difference of opinion as to how much total nitrogen is contained in animal wastes. However, Loehr et al., (1979) estimated the total nitrogen content in various wastes to be about 20-85 mg N / L for untreated waste, 20-50 mg N / L for primary treated waste, and 10-40 mg N / L for secondary treated waste. This is based on annual nitrogen loading to land of 303 kg/ha per year as the low application rate to 3183 kg/ha per year as the high rate application. The study assumed a low rate application of about 5,1 cm/week for a period of 30 weeks whereas the high rate application assumed a 15.2 cm/day for a period of 30 weeks. The manure is applied for 14 consecutive days followed by a 14-day rest period for the mentioned duration (Loehr et al., 1979). This applies for most parts of the U.S. In Canada, Ontario and Quebec are the most conservative provinces in terms of their recommended land application rates with the remaining provinces being less conservative. The recommended swine manure application rate for Alberta and Saskatchewan range from 60 kg/ha and 125 kg/ha to 465 kg/ha and 375 kg/ha respectively, assuming that the manure contains 0.5% N . The actual amount to be applied depends on the soil type, the climatic condition of the area, and the provincial guideline/practice. For example, a 1 tonne manure application in Quebec is assumed to provide approximately 3.5 kg of N whereas the same 1 toime is assumed to provide about 2.5 kg of N in the Maritimes (Agricultural Canada, 1981). These values may underestimate the actual amount of nitrogen contributed by animal waste but may have to be sufficient estimate for the present use. Recommended rates for chemical fertilizer range from 20 kg/ha to 60 kg/ha. The rates may be higher but rarely exceeding 80 kg/ha in non-irrigated croplands of western Canada. Nitrate-N from fertilizer is readily available for plant uptake unlike nitrate-N from animal manure (Agriculture Canada, 1981). Whatever is left is leached out. Very httle nitrogen fertilizer is retained in the soil. In fact, in an earlier experiment conducted in Rothamsted, England, only 2% of the added nitrogen is retained by soil (Jenkinson, 1990, and Goulding, 1990). Although animal manures contain higher concentrations of nitrate-N than chemical fertilizers, the decomposing organic matter tends to slow down the mineralization process thereby lowering the leaching potential (Sutton et al., 1978, and Conrad, 1990). But in order to get the total nitrogen input into the system, other sources such as precipitation and atmospheric nitrogen fixation will have to be considered. Precipitation Input. The annual global input of total nitrogen occurs in both nitrate and ammonical forms. Currently, the actual estimate depends on season climate, method of estimation, and the type of industry activity surrounding sampling locality or location (Loehr et al., 1979). There is great fluctuation in the global estimation of the total nitrogen input to the land surface of the earth in both dry and wet deposition forms. Warneck (1988) gives an estimate of 50 million tonnes per year in the forms of ammonium ion ( N l i ^ ^ ), nitrate ion (NO^"), and nitric acid (HNO3) in precipitation. He also estimated that about 10 million tonnes of ammonium ion (NI^''" ) is deposited in dry form over land and another 16 million tonnes of nitrate ion (NO^") and nitric acid (HNO3) deposited in dry forms over land surfaces of the earth. This brings the total nitrogen deposition by precipitation to about 76 million tonnes per year which is a little bit over the total nitrogen deposit from fertilizers. Burns and Hardy (1975) and Rosswall (1983) gave a range of estimate of the total nitrogen deposition to land as 44 - 200 million tonnes per year. On a regional basis, the total nitrogen input varies from one place to another and also from season to season. Extensive research on the total nitrogen deposition from precipitation has been conducted in Europe. In North-West Europe, total nitrogen dry deposition has been known to be high with estimates ranging from 20 to 60 kg per hectare per year in the heathland (Roelofs et al., 1988). In forests near areas of intensive animal production, total nitrogen dry depositions is found to be more than 100 kg per hectare per year (Ivens et al., 1988). Although the average total nitrogen dry deposition is found to be between 10-20 kg per hectare per year, Derwent et al. (1988) found this to be double the total nitrogen wet deposition from precipitation. In Rothamsted, England, indirect measurements suggest an annual deposition of 35-40 kg of total nitrogen per hectare per year. This value is three to four times higher than the wet deposition. Even though the range of value obtained is by indirect measurement, Goulding (1990) who conducted the actual measurement, reported that there is no significant difference between the two. He concluded that the present total nitrogen deposition on arable land in southern and eastern England is about 40 kg per hectare per year from atmospheric deposition. Research conducted in U.S.A. and Canada on total nitrogen deposition from the atmosphere has not been as extensive, however, with the growing environmental awareness and current growth in atmospheric chemistry research, more data will become available for better total nitrogen deposition estimate from precipitation. Sepp (1971) estimated the total nitrogen input from rain and snow within the United States to vary from 1.1 kg/ha per year to 16.8 kg/ha per year. The actual input depends on the time of the year, geographical location, and the nature of the industries surrounding the geographical location. This estimated input range and the variables controlling the input are applicable to Canada as well for now until a better estimate can be made by atmospheric scientists. Also a better estimate is needed for nitrogen fkation from the atmosphere by soil microorganisms. Soil Nitrogen Fixation. Beringer (1990) encouraged the continual and incremental use of chemical fertilizer if the global food production is going to sustain the rapidly increasing world population. He suggested that biological processes could not "nitrogen-fix" enough of the plant required nitrogen to sustain growth. Loehr et al. (1979), Jenkinson (1990), and Aldrich (1980), however, seem to think that biological contribution of nitrogen is not as insignificant compared to precipitation as Beringer thought. In fact, Jenkinson (1990) estimated the total global input of nitrogen to the land surface of the earth to be about 140 million tonnes per year from biological activities compared to 74 million toimes from fertilizers and 76 million tonnes per year from precipitation (Wameck, 1988). In the U.S.A. , symbiotic nitrogen fixation adds about 3 to 3.6 million metric tons of nitrogen to land per year (Aldrich, 1980). Symbiosis is a process that is mutually beneficial to both plants and microorganisms involved in the relationship. Symbiotic nitrogen fixation can add anywhere from 56 kg/ha per year of nitrogen to 336 kg/ha per year to the soil. Also there has been evidence of nonsymbiotic nitrogen fixation. In fact, nonsymbiotic nitrogen fixation can contribute from 22 to 112 kg/ha per year of nitrogen (Brady, 1974, and Loehr et al., 1979). The precise amount of symbiotic and nonsymbiotic nitrogen contribution is not known exactly, but estimates have shown the contributions to be significant, contrary to Beringer's proposal. However, there is a tendency for a reduction in nitrogen fixation in the presence of other sources of nitrogen such as waste applications (Loehr et al., 1979). As a result, a land receiving waste or fertilizer applications may have a minimal amount of nitrogen input from atmospheric nitrogen fixation. Only in this case can the nitrogen input from nitrogen fixation be considered negligible or insignificant. Since waste is being applied in this case, any form of nitrogen fixation by microorganisms -symbiotic and nonsymbiotic - is considered to be negligible and will not be used in nitrogen loading calculation or total mass balance. However, understanding the activities of soil microorganisms will be crucial in evaluating the transformations that take place between organic and inorganic nitrogen. 2J,.2 Soil Organic-Inorganic Transformations According to Loehr et al. (1979) the amount of nitrogen present in the soil surface layers ranges from about 2240 to 16797 kg/ha. Most of this nitrogen is neither available for plant uptake nor for leaching because it is present in organic forms. However, organic nitrogen undergoes a slow decomposition process with the help of soil microorganisms to an inorganic form in a process called mineralization. The transformation of inorganic-N to organic-N is called immobilization which is the reverse of mineralization. Organic-N * Inorganic-N Immobilization Mineralization encompasses a whole series of reactions of which ammonification and nitrification are two of the most important ones under consideration. Ammonification is the biochemical process of converting organic-N to an ammmonical form such as ammonia, as can be seen from the below equation. microbial R - N H 2 + Ï ^ O y R - O H + NI% + Energy enzyme (ammonia) This process can be carried out by quite a wide variety of heterotrophic microorganisms including bacteria, fungi, and protozoa under favourable conditions. The balance between ammonium release and nitrogen immobilization depends on the carbon-nitrogen ratio of the cell being synthesized, the cell being decomposed, and the overall organism's energy efficiency (Loehr et al., 1979, and Harris, 1988). The ammonium and ammonia are in some form of an equilibrium. But at high p H values, ammonia production predominates. Some of the volatile ammonia is absorbed by vegetation in gaseous form, while the remainder escapes to the atmosphere. In the atmosphere, the ammonia reacts with acidic compounds to form aerosols such as ammonium sulphate and ammonium nitrate which eventually return to the land and sea surfaces as wet and dry depositions. Over the land surface, these aerosol depositions contribute to the total amount of nitrogen present for plant uptake (Whitehead, 1990). However, most soil p H ranges from 4.5 to 7.5. Within this p H range ammonia exists mainly as ammonium ion (NI^"*" ) which can be nitrified in the next step in the transformation process. Nitrification, the biochemical oxidation of ammonium ion to nitrate through an intermediate nitrite production step, is the final phase in the mineralization process. Nitrification occurs primarily under the activity of autotrophic bacteria, but other groups or heterotopic microorganisms can contribute. This is a two step process under well-aerated soils since oxygen availability is very important in ensuring the successful completion of nitrification reactions, as can be seen below (Loehr et al., 1979, and Aldrich, 1980). Nitrosomonas 2 NH4+ + 3 O2 • 2 NO2- + 2 H 2 O + 4 R-" + Energy Nitrobacter 2 NO2' + O2 > 2 NO^' + Energy Under favourable conditions, nitrification can be very rapid. In a field experiment in an Ottawa area clay loam soil, Kowalenko (1978) found that nitrification was very rapid after an addition of N I ^ - N fertilizer, and that within 43 days there were background levels of extractable NI^" ' ' - N . In another study, nitrate-N generation rates of approximately 7 - 25 kg/ha per day were reported after an addition of about 112 kg/ha of ammonium nitrogen fertilizer. The nitrification rate was very rapid (Loehr et al., 1979) confirming the results of the earlier study by Kowalenko (1978). In both cases, an ideal condition was assumed. Under favourable conditions, the obtained result was different. Loehr et al. (1979) observed a nitrite build-up when ammonified wastes were added to highly alkaline soils. The nitrite build-up later resulted in nitrite accumulation in the soils which became toxic to plants. As a result, favourable conditions are prerequisite to nitrification reactions. Reverse of the two mineralization processes - ammonification and nitrification - is immobilization. It is the process of converting inorganic-N to the organic form in microbial or plant tissue. This process makes the nitrogen unavailable to plants and other organisms. Immobilization is highly influenced by the carbon-nitrogen ratio. Addition of waste materials with high C : N ratio (low nitrogen concentration) to soils will result in immobilization initially, however with most of the carbon lost as carbon dioxide with time, mineralization will eventually predominate. Low C : N ratio waste appUcation (high nitrogen concentration) addition to soils, on the other hand, will result in immediate mineralization, as can be seen below (Loehr et al., 1979, and Harris, 1988). Low C : N ratio Immobilization • Mineralization •( High C : N ratio Carbon-nitrogen ratios between or exceeding the range 20 to 25 will result in immediate mineralization with numbers below the range resulting in immobilization first and eventually reversing to mineralization. This supports Kowalenko's findings. After ammonium fertilizer application in an Ottawa area clay loan soil, fertilizer-N immobilization was neghgible in the first 159 days and the amount immobilized subsequently was very small for the rest of the experimental period. While, on the other hand, mineralization in the first two periods sampled averaged 0.77 and 1.10 kg N / h a per day (Kowalenko, 1978). The ammonium fertilizer used by Kowalenko falls under the low C : N ratio with high nitrogen concentration category. Apart from C : N ratio, there are other enviromnental factors that affect the mineralization process, namely temperature, aeration, soil moisture content, and p H (Harris, 1988), Temperature, The optimum temperature for nitrifying bacteria activity ranges from about 25 to 30" C (Harris, 1988) and 27 to 32" C (Brady, 1974). Nitrifying bacteria in areas with higher soil temperatures may, however, adapt to the climatic conditions of the region. This would account for high mineralization rates during the summer months (Kowalenko, 1978), as well as other regional climatic local nitrifying adaptations in both temperate and tropical climates (Harris, 1988; Schmidt, 1982, and Loehr et al., 1979). On the lower temperature extreme, Schmidt (1982) stated that nitrification stops at temperatures below 4 or 5*0, yet nitrification has been reported at below 4°C by a number of people. Fedrick (1956) reported evidence of nitrification at temperatures from 0 to TO, Sabey et al. (1959) detected nitrification at OPC; and Malhi and Nyborg (1986) reported nitrogen mineralization under frozen soils in north-central Alberta. Malhi and Nyborg's observation tend to support Brady's (1974) hypothesis that nitrification occurs down to about (fC and tend to essentially stop at temperatures below approximately -1.1°C. In effect, nitrification occurs over a wide temperature range. However, its occurrence slows down at temperatures below 4 or 5*0. Aeration. Nitrification is an oxidative biochemical process, and as a result, soil aeration status has been speculated to be very important. Although most researchers emphasize well-aerated soil as a requirement, Harris (1988) reported that the rate of nitrification is not seriously affected until the dissolved oxygen concentration falls to about 1.3 x lOr^g/cm? (equivalent to about 0.3% oxygen in solution). Nitrifiers have the ability to recycle nitrate under restricted oxygen conditions. As a result, aeration importance in nitrification processes may have been overestimated. Soil Moisture Content. Both very wet (poorly aerated) and very dry conditions tend to affect the rate of nitrification. A t high moisture contents, the observed effect is primarily due to reduced aeration. In the case of waterlogging, nitrifiers can survive but their activities will be limited to soil-water or air-water interfaces (Harris, 1988). O n the other hand, under extremely dry conditions nitrifiers tend to die off. It looks like nitrifiers are more sensitive to desiccation than they are to waterlogging. Soil p H . According to Harris (1988), there is no clear relationship between p H changes and nitrification rates. However, nitrification does not occur under very acidic conditions except for heterotrophic nitrification in certain cases. In fact, in the p H range 5.5 to 8.0, nothing seems to happen. Any nitrification in a lower p H is due to nitrifiers' adaptation to those conditions (Harris, 1978). On the other hand, Loehr et al. (1979) postulated an optimum p H range of neutral to slightly alkaline condition for the nitrification process to occur. Even though there is no consensus on the effect of p H on nitrification, waste and fertilizer applications are known to affect soil p H . Kowalenko (1987) found a small change in soil p H within one year of fertilizer application; Epstein et al. (1976) found a decrease in p H of one unit after applying sewage sludge and composted sludge to soil samples; and Darusman et al. (1991) found an increase in soil acidification (pH decrease) after twenty years of fertilization with different nitrogen sources. Increases in soil acidification (decrease in soil pH) tend to enhance heavy metal uptake by crops. Crop nitrogen uptake as a method of nitrogen removal is what needs to be enhanced rather than heavy metal uptake. 2.2.3 Nitrogen Removal Mechanisms There are several methods for removing nitrogen from the soil. They are: ammonia volatilization, denitrification, crop harvest, surface runoff, and leaching (Loehr et al., 1979; Whitehead, 1990; and Harris, 1988). Each of the methods will be reviewed as to how essential it is as a nitrogen removal mechanism. Ammonia Volatilization. According to Whitehead (1990), ammonia gas ranks third in abundance following molecular nitrogen (Nj) and nitrous oxide (N2O). Under acidic soils, ammonium ion production predominates. This condition changes as soon as the soil becomes moist and the p H value is increased (greater than 8.5). Ammonium salts tend to undergo some form of chemical reaction resulting in mainly ammonia ( N I ^ ) production. The rate at which ammonia is volatilized depends on the rate of fertilizer or waste application and the soil properties. Research so far has shown that ammonia volatilization is more prevalent in agricultural systems involving animals than from the mainstay arable agriculture (Jenkinson, 1990). If the liquid sludge is surface applied, ammonia loss to the atmosphere may account for about 60 percent or more (Loehr et al., 1979), 14 - 75 percent (Whitehead, 1990), and 20 to 35 million toimes per year worldwide (Warneck, 1988). Ammonia loss from fertilizers is less, ranging from nothing to 53 percent, depending on the type of fertilizer used, soil p H , and soil type. Ammonia volatilization is less in nitrogen fertilizers than from urea. Ammonia volatilization accounts for about 3 to 4% of fertilizer use in England (Whitehead, 1990) and about 18 million tonnes per year globally (Warneck, 1988). These values can be reduced. Injecting or incorporating the sludge/fertilizer into the soil tends to reduce or prevent ammonia volatilization. In addition, soil p H has a major influence on volatilization as does soil type. Research has shown that increasing clay content of the soil reduces volatilization (Loehr et al., 1979). Denitrification. Denitrification is one of the most important causes of nitrogen loss from the soil. Denitrification is the biochemical reduction of nitrate and nitrite to nitrous oxide and molecular nitrogen through the activities of microorganisms mainly heterotrophic bacteria (Harris, 1988, and Loehr et al., 1979). There are two types of denitrification: chemical denitrification (or chemodenitrification) and biological denitrification (or biodenitrification). Since chemodenitrification generally occurs under highly acidic soil condition, this discussion will be limited to biodenitrification. In biodenitrification, which is more applicable to land appUcation of waste, heterotrophic bacteria in the absence of oxygen converts nitrate and nitrite to nitrous oxide and molecular nitrogen under favourable conditions. NO3- y NO2- ^ y N2 To encourage the above reaction, a reliable supply of organic carbon as an energy source is required (Loehr et a l , 1979). According to Harris (1988), nitrate is lost more rapidly from the soil if it is warm, wet and well supplied with organic carbon. This means any loss in carbon content of the soil will ultimately affect the biodenitrification process even under optimum denitrifying conditions. Dorland and Beauchamp (1991) found in their experiment that the denitrification rate at any temperature depends highly on the carbon substrate supply. Lower carbon supply results in higher threshold temperature for denitrification to occur. Temperature plays a major role in controlling the rate of denitrification in the field. According to Harris (1988), an increase in temperature of about ICPC over the range ICPC to 35" C can double the nitrate loss from soil through denitrification. However, denitrification is much reduced in the temperature range 0 to S^C. Other researchers have observed denitrification also within this temperature range. Denitrification has been observed at 3PC (Nommik, 1956), TC (Bremorer and Shaw, 1958), and CfC or lower (-TC) in unfrozen soil (Dorland and Beauchamp, 1991), even though Cho et al. (1979) suggested a threshold temperature of 2.75^0 for denitrification to occur. The effect of soil p H is equally important in determining the rate of denitrification. Under acidic soil (low p H values) only chemodenitrification can occur. Biodenitrification occurs in a neutral to alkaline soil condition. Harris (1988) suggested an optimum p H of 7 to 8 for peak biodenitrification. Finally, the soil moisture content affects the rate of denitrification with high soil moisture content tending to inhibit oxygen availability. Since denitrification thrives under anaerobic conditions, increasing the soil moisture content increases the rate of denitrification. The general consensus indicates that quite a bit of nitrogen is lost through denitrification, but how much is lost is still a subject of much debate. Crop Nitrogen Uptake. Nitrogen uptake by harvestable crops is another important nitrogen removal mechanism. The amount of nitrogen removed depends on a number of variables ranging from crop type, time of the year, and the amount of fertilizer/waste applied to the type of fertilizer used. Increasing the rate of fertilizer application has been known to increase nitrogen uptake in grain and straw (Nuttal and Malhi , 1991), and in ryegrass (Dowdell and Webster, 1980), and (Barraclough et al., 1985). Other researchers have estimated the ryegrass nitrogen uptake to be about 180 kg/ha per year, which is the same as Corn, Coastal Bermidagrass at about 400 kg/ha per year, and wheat at about 60 kg/ha per year (Nuttal et a l , 1991, and Crites et al., 1981). Potato, barley, and soybean nitrogen uptake was estimated to be 230, 70, and 110 kg/ha per year respectively. In effect, significant amounts of nitrogen are removed by crop nitrogen uptake as a nitrogen removal mechanism (Crites et al., 1981) as long as the crops are harvested. Without harvest, the annual nitrogen uptake will still occur but the nitrogen will return to the soil on plant's death as an organic nitrogen. Through transformation, this organic nitrogen may be mineralized resulting in no nitrogen accumulation. However, where transformation process does not occur frequently, nitrogen accumulation may occur over several decades as was observed at Werribee Farm, Melbourne, Australia years ago (Loehr et a l , 1979). Surface Runoff Nitrogen Loss. How much nitrogen is lost from surface runoff varies from site to site, and is still a subject of much debate. With proper management, nitrogen loss from surface runoff can be very minimal, and not constitute a factor in the overall nitrogen loss estimation. Without proper management, however, the amount of nitrogen lost depends on such variables as the amount of natural rainfall, topography of the location, the season of the year, and the application rates/methods (Loehr et a l , 1979). But Aldrich (1988) concluded that nitrogen loss from surface runoff as a nitrogen removal mechanism or source of nitrate pollution to rivers is minimal based on a controlled experiment conducted at Purdue University. Contrary to Aldrich's conclusion, Hubbard et al. (1987) conducted an experiment to evaluate how surface runoff and shallow groundwater quality are affected by centre pivot applied dairy cattle wastes. They concluded that total nitrogen concentration and loads in the surface runoff were highest at high wastewater appUcation rate. A t a nitrogen load of 147.4 kg/ha, the mean total nitrogen concentration was found to be 17 mg/L, which exceeds the recommended limit of 10 mg/L. In effect, all the variables will have to be considered before determining whether nitrogen loss from surface runoff is negligible or not. Soil Nitrate Leaching. Of all the forms of nitrogen, nitrate anion is the one that is washed out of the soil by percolating water. Research has shown that there is both a horizontal and vertical movement of nitrate anion in response to the amount of water moving through the soil. Using components of the water balance equation, as can be seen below, the depth to which nitrate anion can be leached can be roughly estimated (Harris, 1980, and Wild, 1988). As the depth displacement depends on the amount of through-flow water Q = W p - W , - W , , - W t , - A W Q = amount of through-flow water Wp = precipitation (plus irrigation) Wj. = surface runoff Wgy = evaporation from soil surface Wjj = transpiration from crops a W = soil water content increase This is only an indirect estimate. However, several direct measurements have been made on vertical nitrate and on movement in the soil. The conditions that influence nitrate movement in the soil are: • nature of land use activity, • amount of nitrogen supplied and method of application, • the type of nitrogen supplied-ammonium ion versus nitrate ion, • the time of the year applied, • the amount/intensity of the daily rainfall plus irrigation, • the soil type, and • the topography of the location. So far, the measurements indicate nitrate concentrations in the excess of the recommended 10 m g / L in most groundwaters in close proximity to an agricultural operation (Sharpley et a l , 1987; Saffigna and Keeney, 1977; Hubbard and Sheridan, 1983; Hubbard et al., 1985; Hubbard and Sheridan, 1989, and Hubbard et a l , 1991). The general conclusion is that any input of nitrogen from fertilizer/waste application that exceeds the rate of harvestable crop nitrogen uptake and denitrification will result in nitrate leaching in the long run. In fact, Dowdell et al. (1980) suggested that residual effects from fertilizer application can be detected for up to 6 to 9 years or more. How long it takes for residual effects to cease depends on the same conditions that influence nitrate movement in the soil, including the number of years of fertilizer/waste application and the regional climate. Residual effects are of concern over years of fertilizer/waste application. But for the purpose of this research, one of the assumptions is that the contribution from residual effects is negligible. Hence, the nitrate estimation is from the recently applied fertilizer/waste. This assumption simplifies soil nitrogen loading calculation. 2.3 Soil Nitrogen Loading Calculation from Liquid Sludge Although fertilizer and sludge applications to land vary from country to country and from province to province, it is easier for regulatory agencies to recommend an acceptable application rate for fertilizers than it is for animal wastes. This is because animal wastes vary in their nutrient content. For example, total nitrogen content of manures differs from swine to cattle. Similarly, the kind of the treatment the waste undergoes affects the total nitrogen content, with untreated wastes containing the most and wastes subjected to secondary treatment containing the least (Loehr et al., 1979). Since the inorganic nitrogen concentration in liquid sludges or wastes is the limiting factor in land application of wastes, estimating nitrogen mass balance is crucial to minimizing nitrate pollution. Nitrogen mass balance involves an accurate estimation of all nitrogen inputs and the nitrogen losses. Based on the soil nitrogen cycle, the total nitrogen inputs include contribution from the following sources: waste, precipitation, molecular nitrogen fixation, and native nitrogen mineralization. The nitrogen removal mechanisms are: crop nitrogen uptake, loss from leaching/surface runoff, denitrification, ammonia volatilization, and nitrogen accumulated in the soil (Loehr et al., 1979). Nitrogen Inputs = N - output = N - removal mechanisms Waste-N + Precipitation-N + Fixed-N + Mineral-N = Crop-N + Leach-N + Denitrified-N + Volatilized-N + Accumulated-N In order to accurately use the nitrogen mass balance in calculating short-term nitrogen loading, two assumptions are made and they are: 1) That there is no significant addition from nitrogen fixation. 2) There is no accumulation in the soil as either organic-N or inorganic-N (Loehr et al., 1979). These assumptions reduce the nitrogen mass balance equation to: Waste-N + Precipitation-N = Crop-N + Leach-N + Denitrified-N + Volatilization-N The above equation can be used to calculate nitrogen loading rate if all the variables are known for both liquid sludge and wastewater applications. The above equation can also be written as: cP aQ N + = C + — + dN + v N 4.43 4.43 where N = total nitrogen in applied wastes (kg/ha-yr) c = concentration of nitrogen in precipitation (mg/L) P = precipitation (ha-mm/ha-yr) C = crop nitrogen uptake or removal (kg/ha-yr) a = allowable nitrogen concentration in percolating or runoff water (mg/L) Q = volume of water leaving site via percolation or runoff (ha-mm/ha/yr) d = fraction of N which is denitrified (% x 10" ) V = fraction of N which is volatized as ammonia (% x l(r^) 4.43 = conversion factor The volume of water leaving the site via percolation or runoff (Q) depends on precipitation, the water added in liquid sludge, and évapotranspiration. Hence Q = P + zW - E T where P = precipitation (ha-mm/ha-yr) W = liquid sludge addition (ha-mm/ha-yr) E T = potential évapotranspiration (ha-mm/ha-yr) z = moisture fraction in liquid sludge (% x 10"^) Furthermore, the total nitrogen in applied wastes depends on total N fraction in the waste, and if the total solid is known, the total nitrogen in applied wastes (N) becomes: (t)(s)lC^ W N = 4.43 where t = total N fracton in waste (% of solids x 10" )^ s = total solids (% x Iff^) (t)(s)lC^ W Substituting Q = P + zW - E T and N = in the nitrogen mass balance 4.43 equation results in: ts (IC^) W cP a(P + zW - E T ) dts IC^ W vts IC^ W + . — = C + + + 4.43 4.43 4.43 4.43 4.43 Solving the above equation for the liquid sludge addition (W) results in: 4.43C + a(P-ET)-cP W = ts lC^-az-tslC^(d + v) These assumptions are to be made for the use of the above equation for liquid sludge application. They are: 1) Fraction of N volatized as ammonia (v) is zero unless measured experimentally, 2) Fraction of N denitrified (d) is zero unless experimentally measured. 3) The allowable nitrogen concentration in runoff or percolating water is 10 m g / L (the recommend nitrate in drinking water level). 4) The removal of nitrogen in crop or crop nitrogen uptake (C) be set equal to a constant for a given crop for that season. 5) Any excess annual nitrogen is dissolved in the total annual excess of water. 6) Assumptions 4 and 5 may result in nitrogen concentrations leaving the site in excess of the recommended limit of 10 mg/L. With the application of the above assumptions into the final equation, the liquid sludge application (W) equation reduces to: 4.43C + a(P-ET)-cP W = tslO^ - az The use of this equation in calculating acceptable loading rates for liquid sludge should be monitored with care and its performance carefully scrutinized (Loehr et al., 1979). However, chances are that enough conservative assumptions were made that nitrate pollution from the above equation will be minimal. Other calculations use different approaches and a different set of assumptions. Increasingly, mathematical models and computer models are used in nitrate leaching estimation. The accuracy of any model depends on the assumptions made, the parameters used, and the overall use of the model. 2.4 Nitrate Leaching Models For farmers, nitrate loss presents an economic problem in terms of loss in yield and food quality. For the general public, the risk of nitrate being leached to surface and subsurface water bodies spells an enviromnental hazard. Research to date estimates that fertilizer nitrogen recovery by a single harvested crop rarely exceeds 50 to 70% of the applied fertilizer (Hauck, 1973) and sometimes even less. As a result, farmers, agricultural advisers, and environmentalists have a vested interest in understanding fertilizer use efficiency and combatting enviromnental degradation. Models have become a useful tool in predicting environmental problems and developing management practices to deal with them where possible. Where practical solutions are not available, models can suggest an alternative. However, the accuracy or preciseness of any model depends on how good the input data used to construct the model are. This discussion will be limited to three simple models widely used in nitrate leaching prediction. They are: 1) S L I M Model 2) L E A C H N Model 3) SOIL-SOILN Models Slim model was developed originally at the Institute of Arable Crops Research, Rothamsted Experimental Station in England, and has been tested for a number of years (Addiscott, 1977; Lord and Bland, 1991). This model can use detailed soil data where available. Where detailed soil data is not available the model uses topsoil and subsoil texture as input variables. The advantage of this model, apart from its simplicity, is that it uses simple weather data to simulate leaching processes in the field (Lord et a l , 1991). Also the model seems to have room for further improvement. The main disadvantage of the model is its inability to take soil biological processes - mineralization and immobilization - into account. In the experiment conducted by Lord et al. (1991), the land was left barren to avoid biological influences such as mineralization, immobilization, and crop nitrogen uptake. A well-developed model should try to integrate all the input variables in order to give a more representative picture. The L E A C H N Model is a deterministic model. This model is used to estimate water flow and nitrogen movement in an unsaturated soil. The chemical, physical, and biological processes that influence nitrogen availability within the soil profile are described using mathematical functions. The model consists of two main sections: 1) Numerical solutions to the Richard equation for water flow (unsaturated soil condition). 2) The convection-dispersion equation for chemical transport (Hutson and Wagenet, 1991). The final computer program has subroutines that are grouped into different categories that are modifiable independent of one another. The categories are: 1) Water flow 2) Chemical transport 3) Evapotranspiration 4) Heat flow (not coupled with water flow) 5) Rate constant adjustment for temperature and water content. 6) Nitrogen transformations and uptake. 7) Plant growth 8) Plant nitrogen uptake (Hutson et al., 1991). This model accounts for a lot of things that seem to be lacking from the S L I M Model. In terms of being representative, this model seems to depict the actual field conditions. In effect, the L E A C H N Model may be used in making inferences even though it lacks experimentally measured input data. Better estimates from soil survey information will, hopefully, yield better results in the long run. The main disadvantage of the model is its ineptitude in incorporating macropore flow and diffusion in and out of soil aggregates (Hutson et a l , 1991). However, its accounting processes for the chemical, physical, and biological fate of nitrogen, including nitrogen transportation mechanism, is better than the S L I M Model. The SOIL-SOILN Models are used in predicting nitrate leaching rates from arable lands most of the time, although they can also be used in making other predictions. The SOIL-SOILN Models comprise two separate models - SOIL Model and the S O I L N M o d e l The SOIL Model provides the driving variables for the S O I L N Model. The SOIL Model is based on water and heat flow derived from Darcy's and Fourier's Laws respectively. The SOIL Model uses soil and plant properties as well as simple meteorological data such as surface runoff, infiltration, water flow between soil layers, flow to tile drains, soil water content, soil temperature, évapotranspiration, plant water uptake, and hydraulic conductivity function (Bergstrom and Jarvis, 1991). These input parameter values are fed into the S O I L N Model. The S O I L N Model describes the nitrogen inputs from fertilizer/manure and atmospheric deposition, nitrogen transformations such as mineraUzation and immobilization, and nitrogen outputs like nitrate leaching, denitrification, and harvestable crop nitrogen uptake (Bergstrom et a l , 1991). In effect, the S O I L N Model tends to do a nitrogen mass balance for the soil using the input parameters. The usefulness of the SOIL-SOILN Models has been determined in Sweden where according to Bergstrom et al. (1991), large discrepancies were noticed between model predictions and the measured values. They attributed the large error to the ineptness of the models in including macropore flow. Also they suggested that maybe the models did not adequately estimate crop nitrogen uptake, surface runoff and infiltration during snow melt periods. Hence, models are only as good as the data used in their construction. Some models tend to ignore changes in seasonal leaching patterns. This will be reflected in the results obtained using such models. Because in reality, there is a definite pattern in seasonal nitrate leaching. 2.5 Nitrate Seasonal Leaching Patterns Apart from fertilization rates and practices, there seems to be a distinct nitrate leaching variation depending on the season. The variations are, most likely, due to the effects of changing weather conditions such as precipitation and temperature on plant and animal communities. Research to date tends to support the nitrate seasonal leaching pattern hypothesis. In a longitudinal experiment conducted in London, England, by Dowdell et al. (1980), increased amounts of nitrate in drainage water were found following application during the first winter. Only a small increase was observed the subsequent winter. Winter leaching patterns have been well documented in Canada too. Nitrate movement was found to be greatest during the periods of high precipitation and low evaporation rates which correspond to late fall and early spring. High nitrate movement was observed in an Ottawa area clay loam soil during late fall and early spring (Kowalenko, 1978), and in the lower Fraser Valley area of British Columbia (Kowalenko, 1987). In B.C., the greatest risk from nitrate leaching occurs during the fall and winter, with the least risk during late spring and summer months (Kowalenko, 1987). The increase in nitrate leaching potentials in the fall/winter and early spring is due to nitrate availability following end of cropping season, spring snow melt down, mineralized nitrogen from warm spring weather, and high water availability from fall/winter precipitations. During the spring it is much more difficult to determine nitrate loss from the soil. Kowalenko (1987) suggested that nitrate loss in B.C. was due to leaching, contrary to Malhi et al. (1986), Malhi et al. (1986) who conducted field experiments in north-central Alberta. They concluded that the decrease in the amount of nitrate during the early spring was most likely due to denitrification rather than leaching. But without adequate monitoring of both denitrification and leaching, it is hard to tell which one predominates. General consensus is that both removal mechanisms play a major role in determining the spring soil nitrate content. However, if the conditions favourable to denitrification prevails, the amount of nitrate leached out of the soil is greatly reduced. During the growing season (the sununer months), nitrate leaching is minimal. The reasons why nitrate leaching is least during the summer are: • crop nitrogen uptake is high • soil microorganisms are very active • évapotranspiration is very high (less water percolating through the soil). • lower precipitation rate/intensity • Some denitrification can occur according to Kowalenko (1978), thereby further reducing nitrate availability. The summer nitrate loss through leaching and denitrification may be masked by high mineralization rates that occur during the summer months (Kowalenko, 1978). However, an unusually high summer precipitation can result in nitrate leaching as occurred in the summer of 1981. But the general trend is major nitrate leaching between late fall/winter and early spring (Kowalenko, 1987). 2.6 Fertilization Fertilization is very important in increasing or sustaining food yield. But more important is the effect that application rates and time of fertilization have an overall water quahty. Quantity. Research to date shows that nitrate concentrations in the drainage water increase with application rate for a given time period. Regardless of how much research that has been done on application rates, there are few skeptics that tend to dispute the above claim. Vinten, Howard, and Redman (1991) conducted an experiment at Glencorse Mains Farm near Penicuik, Midlothian. Different nitrogen fertilizer inputs were applied to isolated plots up to recommended rate. The results obtained suggested little or no influence of the amount of fertilizer applied on the amount of nitrate leached out of the soil. However, they suggested that accurate nitrogen loading estimation may have been hindered by variable drainflow recovery from the plots, thereby introducing error in the overall experimental analysis. Jenkinson (1991) defined a sustainable agricultural system as one that the yearly organic nitrogen formed roughly balances the amount being decomposed, thereby resulting in no net fertilizer nitrogen accumulation in the long run. A n example of such a system is currently in existence at Rothamsted, England. In the experimental analysis, plots that received 144 kg of nitrogen fertilizer per year since 1843 contained about 3.3 tonnes of organic fertilizer in their topsoil in 1980 with wheat grown each year. Whereas the no -nitrogen fertilizer plots contained 2.9 tonnes of organic fertilizer in their topsoil. Researchers attributed the higher organic nitrogen content in fertilized plots to higher plant debris return, and concluded that nitrogen fertilizer accumulation in soil organic matter is lower in land use for arable agriculture than grazed sward. Hence leaching losses are lower in arable agricultural systems than grazed sward. Darusman et al. (1991) apphed four different types of fertiUzers at different rates ranging from 75 kg/ha to 224 kg/ha per year. The results obtained after 20 years of fertilization show a decrease in soil p H and an increase in nitrate and ammonium concentrations. Hence the potential for nitrate leaching is there. Zwerman et al. (1972) conducted similar experiment in Aurora, New York. They applied moderated (86 kg/ha) to high rate (243.2 kg/ha) of fertilization to 24 plots over a number of years. The results obtained show nitrate concentrations ranging from 3 m g / L to 51.1 m g / L with the higher concentrations corresponding to higher nitrogen fertilizer application rates. Manure applications have been known to follow similar a trend. Liquid swine and cattle manure have similar characteristics. In Lethbridge, Alberta, Sommerfeldt et al. (1988) and Chang et al. (1991) applied cattle feedlot manure to both irrigated and non-irrigated dark brown Chernozemic clay loam soil at varying application rates ranging from 0 to 180 Mg/ha. The results obtained indicate a downward nitrate movement in both irrigated and non-irrigated plots. However, due to low precipitation, high potential evaporation, and the soils' ability to retain water, leaching potential due to percolating water is highly reduced. Nevertheless, if enough water is available (precipitation > 600 mm), there is a tendency for nitrate leaching. This was evident in irrigated plots with higher excess water than non-irrigated plots. In the U.S.A. , Sutton et al. (1978) applied liquid swine waste to Fox silt loam and Chalmers silt loam soils over a period of two years. The application rates range from 45 to 134 metric tons per year with the high rate providing about 378 kg of nitrogen. A separate plot was set up with an inorganic fertilizer application rate of 168 kg/ha per year. The results obtained show that nitrate concentration increased with rate of waste application with little accumulation at the end of the experimental period. Also, that nitrate movement was highest in inorganic fertilizer plots than plot treated with swine waste even though swine waste had the most total nitrogen content. This data supports the generalization that nitrate concentrations increase with application rates until a peak is reached. Apart from application rate, timing is also very important. Time of Application. There are a lot of studies that deal with time of fertilization, nitrate losses and crop yield. Research in Europe, U.S.A. , and Canada has shown nitrate loss potential is increased when fertilizer/waste is applied to frozen or snow covered soil. Thus, winter fertilization is not only an economic waste but also an environmental hazard. Most or all of the applied fertilizer may end up in surface water bodies through surface runoff (Hubbard et al., 1987, and Agriculture Canada, 1981). In fact winter or fall applied fertilizer/waste is more likely to be leached out than that which is applied in the spring or summer. Nuttall et al. (1991) conducted a field experiment in north-eastern Saskatchewan for a period of 3 years. The results of that experiment indicate that autumn applied fertilizer is less effective in encouraging crop yields than spring applied fertilizer. Malhi and Nyborg (1983) attributed the lower autumn efficiency to early spring denitrification caused by snow melt. Nuttall et al. (1991) observed a significant increase in crop yields with spring applied fertilizer than in autumn applied fertilizer, but then suggested that the effectiveness of the applied fertilizer is a function of crop type, soil available nitrogen, and the rate of application. Where the crop response to applied nitrogen is low or a low yielding crop is used, there is no significant difference between autumn applied versus spring applied nitrogen. Campbell et al. (1991) in their experiment noticed that time of fertilizer application only affected crop yield in one location - Swift Current, Saskatchewan. For farmers involved in both winter and spring wheat growing, the researchers suggested that the best time of fertilizer application is mid-October for the Brown soil zone and early spring for the rest of the soil zones. O n the basis of this research, the general conclusion is that yield response to time of application depends on weather, crop type, soil type, and geographical location. Not much research has been done regarding time of application and nitrate leaching in the Lower Fraser Valley region of British Columbia. However, Kowalenko (1987) conducted four field experiments from 1978 to 1982 to determine nitrogen movement in the soil with and without spring or fall applications. The results indicate a low risk of nitrate leaching during the spring and sunraier months due to lower precipitation rate and higher nitrogen crop uptake. Nitrate leaching in this region is highest during the fall and winter months. Thus, after the growing season, any inorganic nitrogen left will be subjected to nitrification and leaching over the fall-winter season thereby making the nitrogen unavailable for spring crop growth. However, in areas of poor drainage, denitrification may predominate. 2.7 Water Table Manipulation Using Drainage Control Proper drainage is crucial to successful farming. In some humid areas, subsurface drainage is installed to help in manipulating the soil moisture conditions within the rootzone. In other situations, the same drainage control can be used to manipulate the water table. Water table, in turn, is used to control the rate of denitrification/nitrification process within the soil profile (Loehr et al., 1979). The conditions that affect both denitrification and nitrification have been reviewed in the previous sections with oxygen concentration playing a very important part in both processes. Accomplishing aerobic or anaerobic conditions within the soil profile depends on the soil type and the efficiency of the drainage control system. For example, in a rapid infiltration soil, enough water has to be added frequently in order to maintain anaerobic conditions within the soil profile (keeping the water table high). Loehr et al. (1979) suggested that meeting the nitrification - denitrification operation requires leaving the water table close to the surface (1.5 to 3.0 m). Water table at this level makes the system manipulation to achieve the desired conditions simple, effective, and less time consuming. In effect, drainage control can be used to skilfully manage nitrate concentrations within the soil profile. How effective drainage control is in nitrate management is still a subject of further research. 2.8 Summary From the literature it is known that the nitrogen fertilizer/waste usage must continue to increase if the world's populations are to be fed properly. With the increase of nitrogen used is the growing concern for nitrogen pollution both to surface and subsurface water bodies. Of all the different forms of nitrogen, nitrate-nitrogen is of the most concern. This is because this mobile anion is easily leached out of soil profile and has been known to cause methaemoglobinaemia - the famous blue baby syndrome. Hence, understanding the soil nitrogen cycle-input, organic-inorganic transformations, and output - is crucial to minimizing nitrate pollution. However, the global nitrogen input seems to outweigh the output mechanisms, as can be seen from Table 2.1 below. In effect accurate nitrogen loading calculation is important in order to avoid exceeding the recommended nitrate in drinking water limit of 10 mg/L. A n d to achieve this, nitrate leaching models, the seasonal leaching patterns, time of application, quantity of waste/fertilizer applied, and drainage control have all been used in the quest for a safer and better drinking water. Table 2.1 Global Nitrogen Input and Output Mechanisms Input Mil l ion tonnes/vear Output Mil l ion tonnes/vear NI% 40 54 N 2 O ~ 12 N O , 36 44 N2 (biological) 140 ~ Fertilizer 74 ~ To Sea ~ 20 Totals 290 130 Source: Jenkinson (1990). ~ Data not available Chapter 3 M A T E R I A L S A N D M E T H O D S 3.1 Experimental Overview The soil samples used in this study were collected from approximately 30 cm beneath the topsoil from a field in Boundary Bay, Delta, B .C . The samples were mixed with composted swine manure at two percent by weight to enhance the drainage ability of this otherwise poorly drained soil. This mixing of compost with soil at two percent by weight was based on field application rate of 66 kg/ha, which was found to be adequate for this type of soil based on an earlier experiment on sludge application (Chieng, 1987). The soil and compost were manually mixed and introduced into clearcast acryhc tubing before fertilization. Also before fertilization, water was slowly added from the top, or introduced from the bottom, to a level of approximately 40 cm height to simulate water table height. Considerable difficulty was experienced in maintaining the water table height at or about the 40 cm height due to problems with trapped air bubbles within the soil plus compost mixture and surface tension. There were also problems from differential saturation. Soil plus compost mixture around the wall of the tube was more saturated than the inner mixtures. After saturation, the soil plus compost mixture was fertilized with different sources of nutrients (liquid chemical fertilizer and liquid swine manure) and also at different rates. Slightly higher application rates were chosen in order to see how the system responds to nutrient over-loading, especially with liquid swine manure. The chemical fertilizer application rate (i.e. 688 ppm N) was slightly higher than the recommended rate by the fertilizer company (Green Valley) for agricultural practices (i.e. 666.7 ppm N). Leachate samples were collected after sitting overnight as day zero. Subsequent samples were collected at two days interval for a period of two weeks. The leachate samples were then analyzed for p H , electrical conductivity (EC), ammonia (NI^ ) , nitrate plus nitrite-nitrogen (NO^ + NO2 " N), and total Kjeldahl nitrogen. It was very difficult to duplicate the same conditions for each of the runs. A detailed description of the runs and their conditions will be made in later sections of this chapter. 3.2 Site and Soil Description The site of the experiment was the Soil and Water Engineering Laboratory at the University of B .C . S k tubes were set up inside the laboratory at room or about room temperature while another six tubes were set up outside the laboratory to simulate field weather conditions. The soil sample used in this study was collected from the Boundary Bay Water Control project site in Delta, British Columbia. The soil is classified as Humic Luvic Gleysol of the Ladner series. The soil textures range from silty clay loam to silt loam. The sample was obtained from approximately 30 cm beneath the topsoil and is composed of mostly silty clay loam with a silt and clay content of about 65.4 and 23,4% respectively. Table 3-1 lists some of the physical and chemical characteristics of soils sampled at this site (Luttnerding, 1981; Driehuyzen, 1983). Table 3-1. Soil Physical and Chemical Characteristics a. Physical Properties Depth Particle Size (%) Texture Bulk Density (cm) sand silt clay *CSSC (g/cm?) 0 - 3 0 3.5 73.3 23.2 sil -3 0 - 5 0 11.2 65.4 23.4 sil 1.40 5 0 - 7 0 21.7 58.8 19.5 sil 1.40 90 - 110 33.1 50.1 16.9 1 1.37 b. Chemical Properties Depth p H p H C Total N C / N (cm) ( H 2 O ) (CaClz) (%) (%) 0 - 3 0 4.4 4.3 5.2 0.421 12.5 3 0 - 5 0 5.6 5.2 1.8 0.135 13.6 5 0 - 7 0 5.8 5.2 0.4 0.035 11.4 90 - 110 4.5 3.9 0.4 0.033 12.2 * Canadian System of Soil Classification (Source: Driehuyzen, 1983). The drainage ability of this acidic soil ranges from moderately poor to poor. This was evident with water table levels within the area. Water table levels are typically at or near the surface for much of the winter in areas with no subsurface drainage. Consequently, the soil was mixed with composted swine manure in the laboratory setup to enhance water drainage. 3.3 Experimental Apparatus Figure 3-1 and 3-2 depict the experimental setup both inside and outside the laboratory. Twelve clearcast acrylic tubes of approximately one meter in length were used for this experiment. The tubes have an inside diameter of 14.8 cm and were sealed at the bottom with a httle hole allowing for water drainage. Six of the tubes were inside the laboratory at room temperature whereas the remaining sk tubes were set up outside the laboratory. The outside tubes were covered with a piece of plywood to prevent rain water from going into the tubes. A fine plastic mesh was placed on each of the tubes to prevent blockade before the soil plus compost mkture was introduced. 3.4 Composted Swine Manure Used In an independent experiment, composted swine manure was mked with poorly drained soil for drainage enhancement and the result was positive. In this study, the same composted swine manure was used for the same purpose. The swine manure used in the composting was collected from a privately owned pig farm located in the Lower Fraser Valley. The swine manure was mked with sawdust at a five to one ratio (on wet weight basis) in order to achieve the proper carbon to nitrogen ratio and moisture content. The compost-sawdust mkture was subjected to an enforced aeration of 0.08 L / m i n . kg of volatile matter. After composting, a particle size analysis and some selected chemical analyses were performed on the composts. Figure 3-3 depicts the particle size distribution analysis for the composted swine manure used in this experiment. The size distributions tend to be between 0,3 mm and 4.5 mm. FIGURE 3 - 1 . SCHEMATIC OF THE EXPERIMENTAL SETUP CLEAR ACRYLIC TUBE lOOMeter 1 T 0,70METERCSOIL + COMPOST) LEACHATE CONTAINER DRAINAGE CONTROL DEVICE FIGURE 3-2. DIMENSIONAL ANALYSIS OF TUBES #1-5. COMPOST BATCH 8 BIN # 2 BIN # 2 (ENFORCED AERATION) 100 % PARTICLES < D 80 60 / 40 • / 20 • I 0 0 1 1 1 1 1 1 1 1 1 r 1 r 1 1 1 10 L d ^ S ZE (mm) FIGURE 3-3, PARTICLE SIZE D ISTRIBUTION Similarly, some selected chemical analysis before mixing the compost with soil revealed a p H of 5.48, a moisture content of about 9.45% after almost two and half years of storage, and a carbon content of 45.26%. 3.5 Liquid Swine Manure and Chemical Fertilizer Used The hquid swine manure used in this experiment was collected from the same location as the one used in composting. Table 3-2 shows some of the characteristics of this manure before application. The manure was covered and stored in a refrigerator until needed for application to minimize nitrification, denitrification, and ammonia volatilization. The chemical fertilizer used was the soluble grade fertilizer 20-20-20 which was purchased from one of the fertilizer distributors in the area. The fertilizer was left in its original form until needed for application at which time an appropriate solution was made. Table 3-2 also shows some of the characteristics of the fertilizer used. Table 3-2. Some of the Characteristics of the Liquid Swine Manure and Chemical Fertilizer Characteristics Liquid Swine Manure Chemical Fertilizer at 200 kg/ha at 400 kg/ha p H 7.30 7.72 7.83 Electrical Conductivity (miscrosiemens/cm) »« 3,500 7,500 Ammonia-N (ppm) 3047.68 454.61 775.02 Nitrate + Nitrite-N (ppm) 1.52 289.21 1038.18 Moisture Content (%) 96.55 * » Total Solids (%) 3.45 * * * Not Applicable ** Greater than 10,000 3.6 Run Procedures Before addition of any nutrients, about 16.2 kg (35.6 lb) of soil and compost mixture was added to tubes #1-5 both inside and outside the laboratory. Tube #6 was the soil alone control. The 16.2 kg (35.6 lb) mass of soil and compost mixture was calculated based on soil plus compost height of about 70 cm (see Appendix A ) , which left enough room for both nutrient and water application. Water was added to each tube to an approximately 40 cm height by either surface application or injection from the bottom to simulate water table height. The tubes were left to sit for between half an hour and an hour to ensure complete saturation. In cases of any drop in the water table height, more water was applied, to bring the level back to about 40 cm. Difficulties were encountered with efforts to keep the water table the same for all the tubes at 40 cm. Problems arose from differential saturation, trapped air bubbles, and surface tension within each tube. These problems resulted in inaccuracy in the volume of water used to keep the water table height at 40 cm. Similarly, different soil mixture contents affected the experimental results too. The drier the soil, the more water and time it took to attain soil and compost saturation. This caused different saturation intensities within and between tubes. 3.6.1 Run 1 The soil and compost used in run #1 had moisture content of about 17.95% and 9.45% respectively (see Table 3-3). In this case, it took less time to saturate the soil plus compost mixture after it was added to the tubes. However, difficulties were still encountered with differential saturation, trapped air bubbles, and surface tension. After saturation and trying to maintain the water table height at 40 cm, tubes #1 and 2 (both with soil and compost) were fertilized with 100 ml of 688 ppm nitrogen and 100 ml of 1376 ppm nitrogen chemical fertilizer respectively. Tubes #3 and 4 (both with soil and compost) were fertilized with 298 ml and 397 ml of liquid swine manure respectively. Tubes #5 (soil and compost alone) and 6 (soil alone) were set up as controls with no nutrient application. The tubes were then left to sit overnight after which about 315.2 ml of water was added to all the tubes to ensure saturation. The 315.2 ml water addition was about four times higher than the calculated water requirement and was based on the average 2-day soil water storage capacity from a previous experiment. This 2-day soil water storage capacity was used to calculate the porosity and the total amount of water required to bring the soil and compost mixture to saturation after fertilization (see Appendix A ) , In some cases, more water was required to achieve the same degree of saturation. Immediately after saturation, leachate samples were collected from each of the tubes both inside and outside the laboratory, and designated as day zero. Subsequent leachate samples were collected every two days for two weeks and all the samples were analyzed for p H , electrical conductivity, nitrate + nitrite-nitrogen, and ammonia-nitrogen. A n attempt to analyze total Kjeldahl nitrogen was unsuccessful due to limited time for mineralization to occur. Table 3-3. Moisture Content of the Soil and Compost before Mixing Material * Moisture Content (% by Weight) Compost 9.45 Soil (old sample) for Run 2 2.53 Soil (still drying) for Run 1 & 2 17,95 Soil (in plastic bags) for Run 1 24.78 * Average of 2 readings. 3.6.2 Run 2 The majority of the soil samples used in run #2 had a moisture content of about 2.53% (see Table 3-3). The compost's moisture content was still the same at 9.45%. After the soil and compost mixture was added to the tubes, attempts to saturate the mixture took longer time. In this case the soil and compost mixture in the tubes were saturated overnight to ensure complete saturation before maintaining the water table height at 40 cm. This overnight saturation resulted in higher soil moisture content in run #2 than run #1, The same difficulties encountered in run #1 were applicable to run #2. The two differences between runs #1 and 2 were higher soil moisture content and higher water table height after fertilization in run #2. Everything else was the same in terms of the length of the run, leachate sample collection, and nutrient application rates. 3.7 Nutrient Application Rates The application of 100 ml of 688 ppm nitrogen and 1376 ppm nitrogen to tubes 1 and 2, equivalent to 200 kg/ha and 400 kg/ha respectively, was based on more than the recommended silage corn field application rate. The 668 ppm nitrogen application to tube #1 was slightly higher than the recommended application rate of 666.67 ppm nitrogen for agricultural practices recommended by the fertilizer company. However, different crops required different nitrogen application rates. In this case, two fertilization rates of 200 kg/ha and 400 kg/ha were chosen for silage corn. The nitrogen calculation for both rates is shown in Appendix A . Tubes 3 and 4 were fertilized with 298 ml and 397 ml of liquid swine manure equivalent to field application rates of 180 tonnes/ha and 240 tormes/ha respectively. The rate calculation for both applications is shown in Appendix A . These apphcation rates of 180 toimes/ha (low) and 240 tonnes/ha (high) were based on past researches with hquid swine manure application. Sutton et al. (1979) applied as much as 134 metric tons/ha of Uquid swine manure to Fox silt loam and Chalmers silt loam soils on their experiment on corn yield. In another experiment, Hanson et al. (1974) applied about 224 metric tons/ha (2% dry matter) and 448 metric tons/ha (wet basis) of liquid swine manure to land and evaluated their effects on corn yield. Since B.C. has no recommended guidelines for swine manure application, researchers have to find suitable application rates with less adverse effects. A l l the nutrients were surface applied and not incorporated. 3.8 Chemical Analysis A l l the leachate samples were collected in polyethylene bottles, stored in a refrigerator, and taken out of the refrigerator a couple of hours prior to any analysis being performed. The p H was measured using Good Digital p H Meter, Model No. 201ATC. The results of the p H measurements were tabulated and are presented in Tables 4-1, 4-2, 4-3 and 4-4 for both inside and outside laboratory set up for runs #1 and #2. The electrical conductivity (EC) measurements were conducted with a portable Myron L PDS Meter, Model E P l l / p H . The meter range was too small to give a reading for the liquid swine manure E C value. The other E C results were tabulated and are given in Tables 4-5, 4-6, 4-7, and 4-8 for runs #1 and 2 for both inside and outside laboratory set up. For the ammonia-nitrogen determination, appropriate standard solutions were made in parts per million (5.0, 7.5, 10.0, 15.0, and 20.0) and used in a Technicon Autoanaylzer II to generate a linear regression curve. The leachate samples were diluted enough that no new dilution was required for the analysis. Appropriate linear regression equation was used to calculate the ammonia-nitrogen concentrations for the samples. The equation was of the form: N H j - N (ppm) = x-coefficient '(Reading) + Constant Both the x-coefficient and the constant were obtained from the generated linear regression curve. The reading was obtained from the Autoanalyzer II print out. The results of the ammonia-nitrogen concentrations were tabulated and given in Tables 4-9, 4-10, 4-11 and 4-12. The nitrate + nitrite-nitrogen measurements were also performed with Technicon Autoanalyzer II but with a different set of standard solutions (0.5, 1.0, 1.5, 2,0, 2.5 and 3.0 ppm). Also, appropriate dilutions were made where applicable to prevent off scale reading. The equation that was obtained where dilutions were made was of the form: NO^ + NO2 - N (ppm) = (x-coefficient * (Reading) + Constant)* dilution factor Just like ammonia-nitrogen, the x-coefficient and the constant were obtained from the linear regression curve, and the reading from the Autoanalyzer II print out. The results of this analysis were tabulated and are presented in Tables 4-13 4-14, 4-15, and 4-16. Finally, leachate sample analysis for total Kjeldahl nitrogen yielded no reading. The leachate samples were too diluted to yield any reading. It is believed that not enough time was allowed for mineralization to occur. Chapter 4 R E S U L T S A N D DISCUSSION Some of the parameters studies in this experiment show some kind of pattern whereas others were not significantly affected by any increase in Uquid swine manure or fertilizer appUcations. In this session, the effects of p H , E C , ammonia-N, nitrate + nitrite-N , T K N , and drainage control on water quality are evaluated. 4.1 pH Measurements The Canadian acceptable range for drinking water p H is 6.5 to 8.5 (Health and Welfare Canada, 1980). Corrosion and encrustation problems occur outside this p H range. However, p H discussion will be limited to water quality and its effects on macronutrients. The p H results from the leachate sample analysis listed in Tables 4-1 to 4-4 for both inside and outside laboratory experiments were variable. The p H results from run #1 for both inside and outside experimental setups were generally below the p H of 6.0, except for tube #6 (soil alone) where the p H values were within the Canadian acceptable drmking water range. For the rest of the tubes in run #1, the p H values were below the p H of 6.0 for most days. This was probably due to the organic matter content of the soil. Organic matter has been known to buffer soil p H in the slightly acid, neutral, and alkaline range. On the other hand, the p H values for run #2 were, in most cases, one p H unit higher than for run #1. This was probably due to a dilution effect resulting from run #2 having considerably higher soil moisture content than run #1. Most of the p H values in run #2 were within the Canadian acceptable drinking water p H range, p H , though, is usually more of a concern for its effect on heavy metal availability than on drinking water quality. The effects of p H on plant nutrients in the soils are more complicated and interrelated. However, some trends do exist in certain cases. Generally, the build up of toxic levels of soluble ions in the soil solution are usually caused by low p H values. This toxic build-up can result in nutritional imbalance. The p H values from run #1 were slightly more acidic than run #2 values, as can be seen from Tables 4-1 to 4-4. This lower p H values in run #1 may affect the macronutrient concentrations under consideration, namely ammonia-N and nitrate + nitrite-N. The detail discussion on how p H affects macronutrients will follow in later sections. However, plants that are able to utilize ammonium forms of nitrogen directly have been known to experience considerable benefit in acidic soils. This is because nitrification (biochemical oxidation of ammonium to nitrate) is slower below the p H of 5.5 (Bohn et al., 1985). 42 Electrical Conductivity (EC) Increasing the soil salt concentration increases the potential for not only salt leaching into both surface and subsurface water bodies but also for salinity effects on plant growth. Earlier measurements of salinity were carried out in terms of the total dissolved solids (TDS), However, this method of determining salinity is less accurate and depends strongly on methodology. Recently, salinity is measured in terms of the electrical conductivity (EC) of the solution. E C measurement is faster and more accurate in establishing salinity problems. Older journals published by Health and Welfare Canada used TDS. Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 5.37 5.12 5.29 5.85 5.25 X Day 2 5.88 5.55 5.66 5.87 6.24 X Day 4 6.33 5.65 5.86 5.53 5.87 X Day 6 5.56 5.04 4.88 5.63 5.62 7.72 Day 8 5.89 5.39 5.32 6.27 5.64 7.96 Day 10 5.67 4.98 4.93 5.06 5.31 7.66 Day 12 5.81 5.42 5.02 5.75 5.50 7.74 Day 14 5.66 4.92 4.65 5.28 5.23 7.74 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Inside Temperature = Room Temperature {ITC to 25.Z'C) X = Indicates inadequate leachate sample for analysis Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 5.72 5.84 5.45 4.76 5.18 6.54 X X Day 2 5.90 5.82 5.42 5.17 5.24 6.85 X X Day 4 5.86 5.82 5.60 5.38 5.21 7.86 X X Day 6 5.36 5.91 5.05 4.69 5.04 6.98 X X Day 8 5.87 6.02 5.33 5.14 5.40 8.01 11.2 Day 10 5.88 5.98 5.11 4.81 5.14 7.99 12.7 Day 12 6.16 5.91 5.20 4.82 5.21 7.05 11.0 Day 14 6.12 5.65 4.78 4.62 5.48 7.05 13.1 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 toimes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 toimes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone X X = Data not available Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 6.18 6.75 6.54 5.96 6.00 6.20 Day 2 6.32 6.57 6.44 6.04 6.05 5.82 Day 4 6.47 6.45 6.43 6.09 6.06 5.95 Day 6 6.93 7.02 6.93 6.53 7.02 6.43 Day 8 6.80 7.37 6.91 6.46 6.74 6.20 Day 10 6.61 6.93 6.51 6.26 6.36 6.05 Day 12 6.81 7.07 6.57 6.20 6.97 5.88 Day 14 6.82 7.20 6.69 6.30 6.94 5.94 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 5.73 5.61 5.54 5.70 5.95 6.17 14.4 Day 2 5.89 5.99 5.98 5.97 6.13 6.13 10.2 Day 4 5.87 6.01 5.99 6.07 6.09 5.94 17.5 Day 6 6.38 6.43 6.36 6.35 6.74 6.40 17.4 Day 8 6.50 6.43 6.11 6.48 6.96 6.44 19.8 Day 10 6.24 6.14 5.89 6.33 6.41 6.10 14.1 Day 12 6.21 6.27 5.91 6.32 7.38 6.33 10.5 Day 14 6.11 6.31 5.78 6.53 7.22 6.04 12.3 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone There is no well documented data on TDS in Canadian drinking water, however reported values from five provinces range from 20 to 3800 ppm. But the Canadian maximum acceptable TDS concentration is set at 500 ppm for drinking water and between 500 - 3500 ppm for irrigation water (Health and Welfare Canada, 1980). The exact suitable TDS concentration for irrigation water depends on the crop type, soil type, drainage abihty of the soil, and site location. The relationship between TDS and E C values (0.1 to 5 mmho/cm) is clear and expressed by the equation below (Bohn et al., 1985): TDS (ppm) = E C (mmho/cm) x 640 However, the higher the E C values, the more difficult it is to establish a clear relationship. There is, though, still a direct relationship between TDS and E C . But for this discussion, E C values will be used. The obvious generalization is that the E C (or the salinity of the soil) increases as the rate of manure or fertilizer application is increased. However, this trend was not observed in this experiment, as can be seen from Tables 4-5 to 4-8. Tables 4-5 and 4-6 show the E C values from run #1 for both inside and outside experimental setups. In both cases, the least E C results were obtained from tube #6 (soil alone). The higher E C values were obtained from outside tube #5 (soil + compost alone). This was probably due to the high salt content of the composted swine manure used which was about 9800 microsiemens/cm from a previous experiment. That same trend was absent from the inside tube #5. Both tubes #2 and #4 for inside and outside laboratory setup for run #1 did not follow the obvious generalization, as can be seen from Tables 4-5 and 4-6. In both cases the rate of manure apphcation did not affect the E C of the soil. Sutton et al. (1978) made a similar observation in their experiment on the effects of liquid swine waste applications to Fox silt loam and Chalmers silt loam soils. In that study, they concluded that soil E C was not affected by increased waste application rate. They gave three reasons for different E C responses to waste applications as follows: - variation in the Na content of the waste applied to soil, - differences in the nutrient holding capacity of the soil, and - finally, differences in the amount of soil water movement within the soil profile. Tables 4-7 and 4-8 present the E C measurements for run #2. The trend on both tables was similar. Increased manure and chemical fertilizer applications did not significantly affect the E C values except for tubes #3 and #4 in Table 4-7. Both tubes, with their soil plus compost nuxture, were fertilized with liquid swine manure at the rates of 180 tonnes/ha and 240 tonnes/ha respectively. Only in this case did increased manure application result in increased E C values from day 0 to day 14. The increase in E C values is not statistically significant enough to support the obvious generalization, however it does follow the expected trend. The lack of trend and consistency in this E C data may be attributed to one or all of the reasons stated above as suggested by Sutton et al. (1978), in addition to lack of complete saturation of the soil plus compost mixture due to air entrapment. Similarly, any reaction that inhibits or affects the presence of ions in solution will affect the E C of the solution. Hence, any complex ion formation that absorbs ions will limit the E C measurements. Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 1900 1620 1400 1000 740 X Day 2 1200 1420 1200 880 640 X Day 4 1000 1300 1000 780 600 X Day 6 990 1100 880 740 540 620 Day 8 990 990 750 730 540 X Day 10 990 930 750 730 540 X Day 12 950 930 750 720 540 570 Day 14 1000 930 740 700 530 570 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Inside Temperature = Room Temperature (22° C to 25.2° C) X = indicates inadequate leachate sample for analysis Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 2100 2300 2000 1100 2200 280 X X Day 2 2100 2000 2000 1100 2000 350 X X Day 4 2000 1900 2100 1150 2100 590 X X Day 6 2000 1800 2000 1150 2000 660 X X Day 8 1950 1750 2000 1100 2000 620 11.2 Day 10 1800 1700 2000 1050 1950 690 12.7 Day 12 1750 1650 1900 1000 1950 680 11.0 Day 14 1700 1590 1900 1000 1900 680 13.1 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone X X = Data not available Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 1400 1000 990 1100 2000 600 Day 2 1400 1000 1000 1200 2000 700 Day 4 1400 1100 1100 1400 2200 790 Day 6 1500 1100 1100 1300 2100 880 Day 8 1500 1150 1300 1300 2000 900 Day 10 1500 1200 1250 1400 1650 930 Day 12 1700 1300 1300 1400 1600 910 Day 14 1500 1300 1300 1300 1450 940 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 990 620 740 400 720 620 14.4 Day 2 1000 850 1000 540 930 550 10.2 Day 4 1000 870 1000 550 990 520 17.5 Day 6 1000 850 1000 590 1000 520 17.4 Day 8 1000 890 950 630 1050 480 19.8 Day 10 1000 900 940 700 1100 480 14.1 Day 12 1000 900 800 700 1100 510 10.5 Day 14 1000 960 860 750 1100 520 12.3 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 toimes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Ammonical nitrogen includes two forms of nitrogen which are ammonium ion (NH^"*" ) and anunonia gas (NHj) . A t p H higher than 8.5, ammonia gas predominates and volatilizes into the atmosphere. A t lower pH, ammonium ion can be held in the soil especially by clay and organic matter. Ammonia gas can dissolve in water up to concentrations of 30 to 40 percent by weight. Both forms in water are not hazardous to humans, although fish and aquatic fauna are extremely sensitive to their concentrations in water bodies. In Canada, there are no recommended guidelines for their presence in water bodies. In the U.S.A. , the suggested upper limit in the water quality criteria is about 1.5 ppm (Aldrich, 1980). Since ammonia-N does not directly affect humans, the rest of this discussion will be focused on its volatilization and transformations. Accordingly to Loehr et al. (1979), close to 60 percent or more of the Uquid sludge manure surface applied volatilizes into the atmosphere as ammonia-N. The rate of volatilization depends on rate and method of manure appUcation, and the properties of the soil. High clay content and slightly acidic soils tend to diminish the rate of ammonia volatilization. How much ammonia-N that volatilized during the period of this experiment was not determined; however, since the p H data in Tabes 4-1 to 4-4 rarely exceeded 8.5, the rate of volatilization wil l be assumed to be insignificant. In this case, it will be safe to assume that the majority of ammonia measured existed in the form of ammonia-N. Tables 4-9 to 4-12 illustrate ammonia-N concentrations in both runs #1 and 2 for outside and inside experimental setups. The only difference between run #1 and #2 was in terms of their respective soil moisture content. The soil moisture content in run #2 was higher than run #1. This may have affected the ammonia-N concentration. Also, there seems to be a difference between the outside and inside ammonia-N concentrations. These differences in concentrations wil l be analyzed a little bit later. A close look at Table 4-9 did not indicate any ammonia-N concentration increase with increase in the chemical fertilizer applied, as can be seen in tubes #1 and #2. However, in tubes #3 and #4, the ammonia-N concentrations were higher in tube #4 than #3. In Tables 4-10, the ammonia-N concentrations seem to have increased with increased chemical fertilizer application. Consistently, the ammonia-N concentrations in tube #2 were higher than in tube #1. This trend was absent with the inside experimental setup in Table 4-9. For the liquid swine manure applications to tubes #3 and #4 in Table 4-10, there was a lack of consistency and pattern unlike tubes #3 and #4 in Table 4-9 (inside the laboratory setup). Similarly, there was a lack of consistency and pattern between tubes #1 and #2, and tubes #3 and #4 in Tables 4-11 and 4-12. Increasing the application rates of chemical fertilizer and liquid swine manure did not significantly affect the ammonia-N concentrations. This may have been due to higher soil moisture content in run #2 than run #1. Also the p H may have been a factor. The p H results in run #2 were generally higher than run #1. Since the p H results in run #2 were approaching neutrality, the tendency for ammonia-N volatilization may have been increased. This will result in low ammonia-N concentrations in run #2 than in run #1. Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 14.41 11.58 10.22 12.27 1.85 0.74 Day 2 14.84 12.56 9.27 9.44 1.02 2.21 Day 4 16.26 13.32 7.22 7.30 1.90 3.73 Day 6 13.65 9.40 8.81 12.56 1.95 3.08 Day 8 14.47 7.87 8.63 12.56 1.27 3.29 Day 10 16.76 3.51 3.04 11.20 2.16 4.30 Day 12 16.37 3.98 3.04 10.05 5.26 2.86 Day 14 13.75 10.04 7.76 10.05 1.85 2.86 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 toimes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 9.90 13.10 2.16 1.95 2.42 1.59 X Day 2 8.74 11.90 2.06 2.06 2.89 3.34 X Day 4 9.72 12.95 2.47 10.45 2.42 4.16 X Day 6 8.96 11.95 2.84 6.02 2.62 2.75 X Day 8 9.11 10.81 2.62 1.75 5.74 2.86 11.2 Day 10 7.09 9.22 1.75 1.99 6.21 2.47 12.7 Day 12 7.22 9.29 2.47 5.21 6.35 2.47 11.0 Day 14 7.54 9.46 1.88 4.82 7.33 2.47 13.1 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone X = indicates data not available Table 4-11. Inside Ammonia-N Concentrations for Run #2 (ppm) Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 5.48 5.41 5.20 4.98 11.41 3.14 Day 2 7.08 8.33 2.32 5.09 9.16 1.83 Day 4 8.98 11.47 6.92 6.06 12.88 1.79 Day 6 14.70 6.28 5.52 9.20 12.49 1.59 Day 8 8.29 6.71 5.48 6.82 10.61 1.61 Day 10 7.36 6.56 5.95 7.69 10.93 1.59 Day 12 8.16 6.71 5.95 8.33 9.63 1.71 Day 14 8.44 6.93 6.26 8.98 10.65 2.30 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 toimes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 2.93 2.36 5.09 4.65 5.85 4.33 14.4 Day 2 4.44 1.79 3.86 3.80 2.88 1.65 10.2 Day 4 4.35 4.06 3.70 2.66 3.01 1.80 17.5 Day 6 3.17 2.85 5.20 4.52 4.11 1.73 17.4 Day 8 2.99 2.68 3.25 2.18 3.22 1.75 19.8 Day 10 2.82 2.58 3.50 2.69 3.01 1.83 14.1 Day 12 3.21 2.99 3.80 2.85 3.33 1.87 10.5 Day 14 3.33 3.17 3.85 2.87 3.56 1.99 12.3 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Even within each run, differences were observed between inside and outside ammonia-N concentrations. Figures 4-1 and 4-1A depict the ammonia-N concentrations for tube #1 for runs #1 and #2. In each run, the inside ammonia-N concentration was higher than the outside ammonia-N concentration. This was also observed in Figures 4-2A, 4-3, 4-4A, and 4-5A. In cases where inside ammonia-N concentrations were higher, this may be attributed to better air circulation outside than inside the laboratory. Any volatilized ammonia outside the laboratory quickly escapes into the atmosphere. Inside the laboratory, volatilized ammonia may not be removed as fast as the outside setup due to poor air circulation within the laboratory. In the other cases, the ammonia-N concentrations were more variable and less consistent. Figures 4-2, 4-3A, 4-4, 4-5, 4-6, and 4-6A show fluctuations in ammonia-N concentrations. The reason for this inconsistency in these figures is not particularly clear, however this may have something to do with the fluctuating p H values, temperature, different soil moisture contents, and differential saturation between and within tubes resulting from air entrapment. The p H values have been known to affect the concentration of ammonia-N obtained. A t higher p H (8.5), ammonia gas predominates and volatilizes into the atmosphere at significant quantities. As a result, tubes with p H values that closely approach neutrality may experience low ammonia-N concentrations due to higher ammonia volatilization. Lower ammonia-N concentrations may also have been due to differences in temperature. A t higher temperatures, ammonium ion conversion to nitrate is elevated. According to Aldrich (1980), the rate of ammonium ion conversion to nitrate depends on temperature. The more ammonium ion that is converted to nitrate, the lower the concentration of ammonia-N. The reverse is true: any decline in soil temperature will slow down ammonium ion conversion and wil l increase the ammonia-N concentration. This is more applicable to the outside laboratory setup than the inside setup. The inside laboratory setup is less susceptible to temperature fluctuations since it was subjected to room temperature. The effects of different soil moisture contents and differential saturation from air entrapment are known, but the extent to which they affect the ammonia-N concentrations in this study are not known. Increased soil moisture content wil l limit anmioniura ion conversion to nitrate as nitrification requires well aerated soil. This will probably lead to ammonium ion build-up in the soil, thereby increasing ammonia-N concentrations. Similarly, incomplete saturation will lower the concentrations of anmionia-N, thereby introducing error in the result analysis. The effect of individual variable (pH, temperature, soil moisture content, and differential saturation) on ammonia-N has been explained above, however the combined effect of all the variables is difficult to explain. This is evident in some of the figures obtained. For example, the irregularities in Figures 4-3A and 4-4 can not be explained by either one of the above variables mentioned or by all of the variables combined together. This introduces an element of uncertainty in the result analysis. However, pattern or no pattern, tubes #5 and #6 without any fertilizer or liquid swine manure had the least ammonia-N concentrations. AMMONIA-N CONCENTRATION!ppm) 6 -4 -2 -QI 1 1 1 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY B DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC, - + - OUTSIDE CONC. FIGURE 4-1.AMMONIA-N CONCENTRATIONS FOR TUBE #1 RUN #1 INSIDE AND OUTSIDE AN/MONIA-N CONCENTRATIONS FOR TUBE #1 RUN # 2 AMMONIA-N CONCENTRATIONIppm) Q l 1 1 1 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY B DAY B DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. - ^ - OUTSIDE CONC. AMMONIA-N CXDNCENTRATIONIppm) g I r 1 1 ] 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY B DAY ID DAY 12 DAY H TIME œNCENTRATION — INSIDE CONC, OUTSIDE CONC. FIGURE 4-2,AMMONIA-N CONCENTRATIONS FOR TUBE #2 RUN #1 INSIDE AND OUTSIDE AMMONIA-N CONCENTRATIONS FOR TUBE #2 RUN #2 AMMONIA-N CONCENTRATION(ppm) DAY 0 DAY 2 DAY 4 DAY B DAY 8 DAY ID DAY 12 DAY 14 TIME CONCENTRATION INSIDE CONC. OUTSIDE CONC. AMMONIA-N CONCENTRATIONIppm) Q l 1 1 1 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC, - t - OUTSIDE CONC. FIGURE 4-3.AMMONIA-N CONCENTRATIONS FOR TUBE #3 RUN #1 INSIDE AND OUTSIDE AMMONIA-N CONCENTRATIONS FOR TUBE #3 RUN #2 AMMONIA-N CONCENTRATION(ppm) 2 -1 -g l 1 1 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC, - t - OUTSIDE CONC, AMMONIA-N CONCENTRATION!ppm) DAY 0 DAY 2 DAY A DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. - t - OUTSIDE CONC. FIGURE 4-4.AMMONIA-N CONCENTRATIONS FOR TUBE #4 RUN #1 INSIDE AND OUTSIDE AMMONIA-N CONCENTRATIONS FOR TUBE # 4 RUN #2 AMMONIA-N CONCENTRATION!ppm) 0 I 1 1 1 1 r 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. OUTSIDE CONC. AMMONIA-N œNCENTRATION(ppm) 81 g l 1 1 1 1 \ 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY ID DAY 12 DAY 14 TIME CONCENTRATION NSIDE CONC. OUTSIDE CONC. FIGURE 4-5.AMMONIA-N CONCENTRATIONS FOR TUBE #5 RUN #1 INSIDE AND OUTSIDE AMMONIA-N CONCENTRATIONS FOR TUBE #5 RUN #2 ANA40NIA-N CONCENTRATION!ppm) 2 -g I I I I I I I I DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. OUTSIDE CONC. AMMONIA-N CONCENTRATION(ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. OUTSIDE CONC. FIGURE 4-6,AMMONIA-N CONCENTRATIONS FOR TUBE #6 RUN #1 INSIDE AND OUTSIDE AMMONIA-N CONCENTRATICWS FOR TUBE #6 RUN #2 , AMN/DNIA-N CONCENTRATION(ppm) 1 -0 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION INSIDE CONC. - ^ O U T S I D E CONC. With all the inconsistencies in the tubes fertilized with chemical fertilizer and hquid swine manure in both outside and inside laboratory setups, tubes #5 (soil + compost alone) and tubes #6 (soil alone) exhibited more consistency with ammonia-N concentrations, as can be seen from Figures 4-7, 4-7A, 4-8, and 4-8A. This was probably because the ammonia-N concentrations were so low, due to lack of fertilizer or swine manure application, that the variables (pH, temperature, soil moisture content and differential saturation) did not play much of a part in altering the concentrations from day to day. 4.4 Nitrate + Nitrite-Nitrogen Analysis Unlike the ammonia-N concentrations in the runs, the nitrate + nitrite-N concentrations show a decreasing trend as the experiment approaches day 14. The concentrations never quite approached zero as indicated in some of the figures, however the concentrations were below the lowest standard detectable limits of 0.25 and 0.10 ppm in some cases, as can be seen from Tables 4-13 to 4-16. Nitrate-N concentration in water bodies, unlike ammonia-N, does not directly affect the fish and aquatic fauna, however it can lead to accelerated eutrophication in water bodies. Nitrate-N in conjunction with phosphate in water wil l result in increased plant activities which wil l eventually deplete the dissolved oxygen concentration of water. The immediate result is fish ki l l and a decrease in algal activities; it is also objectionable from an aesthetic standpoint. But the most serious effect is the high biological oxygen demand generated by the decaying plant matter. This leads to a change in the water trophic status. There have been documented cases of water bodies, especially temperate lakes, going from eutrophic to an oligotrophic status due to nitrate-N concentration changes within the water body. Lake Erie and Lake Ontario, for example, have been known to have undergone INSIDE AMMONIA-N CONCENTRATIONS FOR FOR TUBES #1-6 RUN #1 AMMONIA-N CXDNCENTRATIONlppm] — + " ' jr.?-*. ., , 1, 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME TUBE # — TUBE # 1 TUBE tt2 TUBE # 3 - a - TUBE # 4 TUBE # 5 - » - TUBE # 6 FIGURE 4-7.AMMONIA-N CONCENTRATIONS FOR TUBES #1-6 RUN #1 INSIDE AMvlONIA-N CONCENTRATIONS FOR FOR TUBES #1-6 RUN # 2 AMMONIA-N CONCENTRATION!ppm) g l ^ 1 1 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY ID DAY 12 DAY 14 TIME TUBE » TUBE #1 - t - TUBE # 2 TUBE # 3 TUBE # 4 - » * - TUBE # 5 - » - TUBE # 6 CDUTSIDE AMMONIA-N CONCENTRATIONS FOR FOR TUBES #1-6 RUN #1 AMMONIA-N CONCENTRATIONIppm) Q l 1 > 1 < 1 , 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME TUBE # TUBE « \ TUBE # 2 ^ TUBE # 3 -e- TUBE « 4 - X - TUBE # 5 - » - TUBE # 6 FIGURE 4-B.AMMONIA-N CONCENTRATIONS FOR TUBES #1-6 RUN #1 OUTSIDE AMMONIA-N CXDNCENTRATIONS FOR FOR TUBES #1-6 RUN #2 AMMONIA-N CONCENTRATION(ppm) .^^^^^^^ 1 1 J 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME TUBE # — TUBE » t - * - TUBE »2 - • - TUBE # 3 - a - TUBE # 4 - X - TUBE «5 TUBE «6 changes due to nitrate loading from nonpoint sources (Bertrand et al., 1989). Nitrate-N in the drinking water has been alleged to cause several human health problems ranging from birth defects to nervous system impairment. However the only verified and well documented effect has been methaemoglobinaemia. This disease affects human infants because of the nitrate in the drinking water being converted to nitrite. Nitrite when absorbed in the bloodstream inhibits the blood's ability to absorb and carry oxygen. This has led to the recommendation that the maximum acceptable nitrate-N and nitrite-N concentrations in the drinking water be limited to 10 ppm and 1.0 ppm respectively (Health and Welfare Canada, 1980). Nitrite, though, is very unstable and is readily converted to nitrate. Where both nitrate and nitrite-N are present, the combined nitrogen forms in the drinking water should also not exceed 10 ppm (Health and Welfare Canada, 1980). Nitrate + nitrite-N concentrations on day 14 of the experiment were considerably lower than their initial concentrations on day 0 in both runs, as can be seen from Tables 4-13 to 4-16. In all the runs and each of the tubes, there was a decreasing nitrate + nitrate-N concentration trend with time. In both runs, there was a more drastic decrease in nitrate + nitrite-N concentration with time with the inside experimental setup than the outside one. Unlike the ammonia-N data, the nitrate + nitrite-N measurements show a more consistent concentration decreasing pattern. The rate of the decrease varies from tube to tube and also between runs. The differences in the rate of concentration decline can be attributed to differences in p H , temperature, soil moisture content, and nutrient availability. A close look at nitrate + nitrite-N concentration in each tube between outside and inside experimental setup will help to explain presence or absence of a pattern. Figures 4-9 and 4-9A depict the nitrate + nitrite-N concentrations for inside and outside experimental setups for both runs #1 and #2. In each of the figures, the outside nitrate + nitrite-N concentration was higher than the inside. In each case, there was a decreasing concentration trend with time, however the rate of decline was more drastic in run #2, as can be seen from Figure 4-9A. This was probably due to higher soil moisture content in run #2 than run #1 among other things. This can also explain, in part, the differences in concentrations between inside and outside experimental setups. Figures 4-10 and 4-lOA show a different pattern but a similar trend to that seen in Figures 4-9 and 4-9A. The differences in pattern can be attributed to differences in soil moisture content within and between runs. The similarity in trends can be attributed to the tubes undergoing the same processes. The change that occurs seems to be more pronounced between day 4 and day 8. This change is a result of denitrification, defined as the biochemical nitrate and nitrite reduction by microorganisms to gaseous nitrogen and oxides of nitrogen (Loehr et al., 1979). The rate of denitrification is affected by p H , temperature, soil moisture content, and nutrient availability. Denitrification proceeds in some tubes at a faster rate than others. In order to understand the differences in the rate of denitrification, the conditions that limit or favour this process are carefully evaluated in a subsequent section. Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 102.43 204.17 156.13 66.10 54.58 8.37 Day 2 94.62 280.92 68.25 22.28 19.35 0.45 Day 4 11.05 418.40 55.56 49.47 12.35 0.45 Day 6 0.51 38.39 19.82 0.48 0.45 0.45 Day 8 0.45 9.11 0.45 0.45 0.45 0.45 Day 10 0.45 0.45 0.45 0.45 0.45 0.45 Day 12 0.45 0.45 0.45 0.45 0.45 0.45 Day 14 0.43 0.43 0.43 0.43 0.43 0.43 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Table 4-14. Outside Nitrate + Nitrite-N Concentrations for Run #1 (ppm) Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 228.19 260.41 219.99 90.71 252.21 16.01 X Day 2 252.21 196.55 390.38 112.58 236.97 8.07 X Day 4 192.46 143.24 189.14 72.16 246.94 8.31 X Day 6 171.37 133.91 198.13 100.19 227.52 8.89 X Day 8 175.12 151.14 201.51 90.82 224.14 12.95 11.2 Day 10 142.34 134.66 176.05 134.15 188.23 9.52 12.7 Day 12 154.95 155.51 200.76 93.15 186.96 11.43 11.0 Day 14 131.78 126.26 211.80 112.46 195.24 11.46 13.1 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone X = indicates data not available Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 Day 0 75.57 30.91 59.67 98.42 218.63 1.18 Day 2 37.00 0.66 23.19 69.21 208.46 0.66 Day 4 0.66 0.66 0.66 59.67 198.76 0.66 Day 6 0.66 0.66 0.66 10.21 203.77 0.66 Day 8 0.66 0.66 0.66 0.66 1.59 0.66 Day 10 0.66 0.66 0.66 0.66 1.20 0.66 Day 12 0.66 0.66 0.66 0.66 0.93 0.66 Day 14 0.04 0.04 0.04 0.04 0.36 0.04 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone Table 4-16. Outside Nitrate + Nitrite-N Concentrations for Run #2 (ppm) Tube #1 Tube #2 Tube #3 Tube #4 Tube #5 Tube #6 A i r Temp. °C Day 0 72.59 26.77 63.25 23.13 57.69 32.96 14.4 Day 2 57.09 25.84 71.00 22.79 67.62 30.64 10.2 Day 4 56.69 18.89 71.00 13.75 69.01 27.00 17.5 Day 6 32.80 0.98 113.37 0.91 88.53 2.35 17.4 Day 8 1.18 0.66 1.21 0.66 1.21 0.66 19.8 Day 10 0.81 0.66 0.85 0.66 0.91 0.66 14.1 Day 12 0.30 0.04 0.37 0.04 0.54 0.04 10.5 Day 14 0.10 0.04 0.21 0.04 0.25 0.04 12.3 Tube #1 = Soil + 2% Compost + Chemical Fertilizer at 200 kg/ha Tube #2 = Soil + 2% Compost + Chemical Fertilizer at 400 kg/ha Tube #3 = Soil + 2% Compost + Liquid Swine Manure at 180 tonnes/ha Tube #4 = Soil + 2% Compost + Liquid Swine Manure at 240 tonnes/ha Tube #5 = Soil + 2% Compost Alone Tube #6 = Soil Alone The rest of the figures (Figures 4-11 and 4-11A to Figures 4-15 and 4-15A) all exhibited different patterns but the similarities were the same in terms of a decreasing concentration trend. Figures 4-15 and 4-15A depict inside nitrate + nitrite-N concentrations for tubes #1 - 6 for runs #1 and #2. Consistently, tube #6 (soil alone) had the lowest nitrate + nitrite-N in both runs and consequently showed the least change. On the other, tube #5 (soil + compost alone) had the highest nitrate + nitrite-N concentrations in both runs, and hence underwent the most change in terms of denitrification. The nitrate + nitrite-N concentration in other tubes were subjected to denitrification at different rates. Based on this result, the rate of manure and fertilizer applications to soil did not significantly influence the nitrate + nitrite-N concentrations in the inside experimental setup for both runs. Nitrate + nitrite-N concentrations were consistently higher in tube #5 (soil + compost alone) than tubes #2 (soil + compost + chemical fertilizer at 400 kg/ha) and #4 (soil + compost + liquid swine manure at 240 tonnes/ha). Both the chemical fertilizer and Uquid swine manure applications exhibit a "hindering or suppressing effect" on nitrate + nitrite-N concentrations within the tubes. This tends to contradict the popular beUef that nitrate + nitrite-N concentrations increase with increasing chemical fertilizer or liquid swine manure application. The inside laboratory room temperature is higher than the field condition except during the sunmier months. Also, the inside laboratory room temperature is subjected to less fluctuation. There is lower air circulation in the laboratory and probably higher soil moisture content especially for run #2. The application of composted swine manure to enhance drainage ability of this poorly drained soil ended up providing sufficient carbon supply source for denitrification to occur. These conditions and deviations from the widely practiced procedure probably explain the above mentioned contradiction. The same pattern and similarity in decreasing NO3 + NO2 -N concentrations were observed for tubes #1-6 run #2 (outside setup) in Table 4-16A. The similarity was less obvious in Figure 4-16 for run #1. This was probably because difficulties were encountered in achieving a complete saturation in run #1 and this probably resulted in low soil moisture content. 4.4.1 Factors that Affected Nitrate + Nitrite-N Concentrations Under field conditions, nitrogen is lost through ammonia volatilization, denitrification, crop harvest, surface runoff, and leaching. In this experiment, some of these conditions never applied. The ammonia-N was accounted for except for the volatilized ammonia. The crop uptake is not applicable since no crop was grown. There was no loss through surface runoff since the tubes were not flooded to overflow. Hence this limits nitrate + nitrite-N losses to leaching and denitrification. The leached NO^ + N O ^ - N was collected and analyzed. The results of the analysis have been presented in the preceding section. However, the denitrified nitrogen loss cannot be accounted for since denitrification was not measured. But since everything else has been accounted for or is not applicable, inferences can be made as to the unaccounted nitrogen loss. In this case, it is safe to assume that any NO^ + N O j - N that was not leached out has been denitrified. NITRATE * NITRITE-N CONCENTRATIONS(ppm) 3001 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME œNCENTRATION — INSIDE CONC. OUTSIDE CONC. FIGURE 4-9.NITRATE + NITRITE-N CONCENTRATIONS FOR TUBE #1 INSIDE AND OUTSIDE (NITRATE + NiTRITE]-N CONCENTRATIONS FOR TUBE #1 RUN # 2 NITRATE t NITRITE-N CONCENTRATIONS(ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION INSIDE CONC. OUTSIDE CCMC. NITRATE * NITRITE-N CX)NCENTRATIONS|ppm) 5001 DAY 0 DAY 2 DAY A DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. - t - OUTSIDE CONC. FIGURE 4-10.NITRATE + NITRITE-N CONCENTRATIONS FOR TUBE #2 INSIDE AND OUTSIDE [NITRATE + NITRITE)-N CONCENTRATIONS FOR TUBE #2 RUN #2 NITRATE t NITRITE-N CONCENTRATIONS(ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. - t - OUTSIDE CONC. NITRATE t NITRITE-N œNCENTRATIONS]ppm) DAY D DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC, - + - OUTSIDE CCMC, FIGURE 4-11.NITRATE * NITRITE-N CONCENTRATIONS FOR TUBE #3 INSIDE AND OUTSIDE (NITRATE + NITRITE)-N CONCENTRATIONS FOR TUBE #3 RUN #2 NITRATE + NITRITE-N CONCENTRATIONS!ppm) 1201 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC, - + - OUTSIDE CCMC. NITRATE t NITRITE-N œNCENTRATIONS|ppm) g I 1 1 1 1 1 1 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME œNCENTRATION — INSIDE CONC. - ^ - OUTSIDE CONC. FIGURE 4-12.NITRATE » NITRITE-N CONCENTRATIONS FOR TUBE #4 INSIDE AND OUTSIDE (NITRATE + NITRITE)-N CONCENTRATIONS FOR TUBE #4 RUN #2 NITRATE * NITRITE-N CONCENTRATIONS(ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. - t - OUTSIDE CONC. 300 250 200 150 100 50 NITRATE + NITRITE-N œNCENTRATIONS|ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 TIME DAY 10 DAY 12 DAY 14 CONCENTRATION INSIDE CONC. OUTSIDE CONC. FIGURE 4-13.NITRATE * NITRITE-N CONCENTRATIONS FOR TUBE #5 INSIDE AND OUTSIDE (NITRATE + NITRITE)-N CONCENTRATIONS FOR TUBE #5 RUN #2 NITRATE + NITRITE-N COMCENTRATIONSlppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 AY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. - + - OUTSIDE CONC. NITRATE * NITRITE-N CONCENTRATIONS(ppm) 5 g I ' DAY 0 DAY 2 DAY A DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION INSIDE CONC. - t - OUTSIDE CONC. FIGURE 4-14.NITRATE * NITRITE-N CONCENTRATIONS FOR TUBE #6 INSIDE AND OUTSIDE (NITRATE + NITRITE)-N CONCENTRATIONS FOR TUBE #6 RUN # 2 NITRATE + NITRITE-N CONCENTRATIONS(ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME CONCENTRATION — INSIDE CONC. OUTSIDE CONC. INSIDE NITRATE + NITRITE-N CONCENTRATIONS FOR TUBES #1-6 RUN #1 NITRATE * NITRITE-N CXDNCENTRATIONS|ppm) 5001 DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 DAY 10 DAY 12 DAY 14 TIME TUBE # — TUBE #1 - 1 - TUBE # 2 TUBE # 3 - a - TUBE # 4 TUBE # 5 - » - TUBE # 6 FIGURE 4-15,NITRATE * NITRITE-N CONCENTRATIONS FOR TUBES 1-6 INSIDE NITRATE + NITRITE-N CONCENTRATIONS FOR TUBES #1-6 RUN #2 NITRATE t NITRITE-N CONCENTRATIONS(ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY TIME DAY 10 DAY 12 DAY 14 TUBE ff — TUBE #1 - * - TUBE # 2 - * - TUBE # 3 - » - TUBE # 4 - • ^ TUBE # 5 - » - TUBE # 6 dJTSIDE NITRATE + NITRITE-N CONCENTRATIONS FOR TUBES #1-6 RUN #1 NITRATE t NITRITE-N CXDNCENTRATIONS!ppm) 5001 0^ ' DAY 0 DAY 2 DAY 4 DAY B DAY 8 DAY 10 DAY 12 DAY 14 TIME TUBE ff TUBE * 1 TUBE # 2 TUBE # 3 - e - TUBE # 4 TUBE # 5 - » - TUBE # 5 FIGURE 4-I6.NITRATE * NITRITE-N CONCENTRATIONS FOR TUBES 1-B OUTSIDE NITRATE + NITRITE-N CONCENTRATIONS FOR TUBES #1-6 RUN #2 NITRATE * NITRITE-N CONCENTRATIONS|ppm) DAY 0 DAY 2 DAY 4 DAY 6 DAY 8 TIME DAY 10 DAY 12 DAY 14 TUBE # TUBE #1 TUBE « 2 TUBE # 3 TUBE # 4 TUBE # 5 TUBE *f6 Nitrate + nitrite-N loss through denitrification can be quite extensive within a short period of time if the conditions are favourable for the process to proceed, as can be seen in Figures 4-15, 4-15A, 4-16, and 4-16A. The conditions or factors that affected N O j + N O 2 - N concentrations in this experiment were: • p H • temperature • nutrient availability (carbon source) • soil moisture content Soil p H is important in determining the rate of denitrification. In highly acidic soil (low p H value), chemodenitrification is predominant. This is only of minor significance in waste application to land. Biodenitrification played the greatest part in this experiment, since the p H of soil plus compost mixture with the appropriate nutrient application goes from slightly acidic to neutrality. This trend conforms with Harris' (1988) prediction that biodenitrification peaks at an optimum p H of about 7 to 8. This was probably why the NO3 + N O j - N concentrations were generally low in the run #2. The p H values were, in most cases, higher by as much as one p H unit, as can be seen from Tables 4-3 and 4-4. Although denitrification has been observed at 3^C (Nommik, 1956), TC (Bremner et al., 1958), and (fC or lower (-2°C) in unfrozen soil (Dorland et a l , 1991), higher temperature always results in greater rate of denitrification. The inside experimental setup was subjected to room temperature which varied between 22'C to as high as 25S C on a sunny day. This according to Harris (1988), puts the temperature range within the optimum nitrate loss from the soil through denitrification. The outside experimental setup, on the other hand, was subjected to daily night and day temperature fluctuations. A t night the temperature dropped as low as 3 to S'C. Whereas during the day, measurements between 1200 and 1400 hours ranged from a low of 10.2°C to a high of 19.8f C, as can be seen from Tables 4-14 and 4-16. This daily temperature fluctuations probably affected the rate of denitrification outside the laboratory. This explains why the NO^ + N02"N concentrations were, for the most part, consistently higher in the outside setup than in the inside setup. Nutrient availability (or carbon supply source) is another factor that probably influenced the NO3 + N O ^ - N concentrations in this experiment. For denitrification to occur, a reUable organic carbon supply as an energy source is required by the heterotrophic bacteria. This organic carbon supply source was provided when composted swine manure was added to the soil. The composted swine manure not only enhances the drainage ability of the soil but also provides the energy supply required by these bacteria to facilitate denitrification. If the temperature is right, the process proceeds at a fast rate. According to Dorland et al. (1991), the rate of denitrification at any temperature depends heavily on the organic carbon substrate supply. Since the organic carbon supply source was there in conjunction with the appropriate temperature, the conditions needed for denitrification (or biodenitrification) were met. This also helps in explaining the lower NO^ + NO2-N concentrations with the inside experimental setup. Finally, for biodenitrification to occur, the soil moisture content has to be high. This inhibits oxygen availability thereby creating the required anaerobic conditions. Under anaerobic conditions, bacteria use the chemically combined oxygen in nitrate to facilitate the biodenitrification process (Loehr at el., 1979). This is another factor that was met with the experimental procedure. The anaerobic conditions due to high soil moisture content or high water table were probably better achieved in run #2 than in run #1. Run #2 was saturated overnight (due to low soil moisture content) to ensure complete saturation. This not only increased the soil moisture content but also raised the water table to a higher level than what was observed in run #1. This is probably the other explanation for the lower NO^ + N O 2 - N concentrations in run #2. These factors and their corresponding effects conform with one of Harris' (1988) major observations which states that nitrate loss is more rapid from the soil that is warm, wet and well supplied with organic carbon. This happens to be very applicable in this experimental setup. 4.5 Total Kjeldahl Nitrogen (TKN) The T K N concentrations in the leachate samples were undetectable. This was because the T K N , after undergoing the minimum dilution to prevent acid interference and equipment corrosion, has been diluted too low to be detected by the Autoanalyzer. Similarly, the two week experimental time period does not allow enough time for the organic-N to be mineralized. More time is probably required for any mineralization process to occur. 4.6 Drainage Control Effects on Nitrate + Nitrite-N Even though difficulties were encountered in an effort to accurately control the water table level due to air entrapment and surface tension, considerable success was achieved in observing the effects of water table height on denitrification. The water table height was higher in run #2 than run #1. Because of the overnight saturation in run #2, the 315.2 ml of water raised the water table higher than what was observed in run #1. This led to complete inundation of the soil plus compost mixture with water in run #2 and ensured the anaerobic conditions required for biodenitrification to occur. The soil plus compost mixture was not completely inundated with water in run #1. This was probably why the N O j + N O ^ - N concentrations were higher in run #1. Hence, restricted or controlled drainage can be used to manipulate the water table and hence the rate of nitrification - denitrification process. Chapter 5 C O N C L U S I O N These experiments were designed to study the effects of different nutrient apphcations and drainage control on water quality. The study itself generated a lot of data on p H , E C , ammonia-N, and nitrate + nitrite-N. Based on the result analysis, it is clear that each run is different from the subsequent run due to variation in soil moisture content, water table height, and temperature (especially for the outside setup). As a result, data is different from run to run. However, certain deductions can be made. Based on the results, there was no significant difference between liquid swine manure appUcation as opposed to chemical fertilizer appUcation in terms of NO3 + NO2 -N concentrations on day 14 of the experiment. In fact, the NO3 + NO2 -N concentrations on day 14 were weU below the recommended safe drinking water level of 10 ppm for aU the runs except for the outside tubes in run #1. The consistently high NO3 + NO2-N concentration in all the outside tubes in run #1 was probably due to inadequate saturation within the soil + compost mixture. Similarly, the outside temperature for run #1 was on the average 2.5°C lower than the outside temperature for run #2. Generally under normal circumstances, any increase in manure or chemical fertilizer application increases the potential for NO3 + NO2-N leaching. This trend was not observed in the present study. Probably because the application of composted swine manure to enhance the drainage ability of the poorly drained soil did more than drainage enhancement. The compost application may have supplied the organic carbon required by heterotrophic bacteria to facihtate biodenitrification. High manure or chemical fertilizer application did not result in any increase in NO3 + NO2 -N leaching potentials in this study. Contrary to the initial belief about nutrients being deficient in composted swine manure, tube #5 (soil + compost alone) consistently had about the same or in some cases as high NO3 + NO2 -N concentrations as the tubes fertilized with extra nutrients. Hence, composted swine manure could have been used in drainage enhancement as well as supplying the plant required nutrients. The other generated results were more variable and less consistent with trend. The p H results had no definite trend except that run #1 p H values were lower than run #2 values. This may have been due to higher soil moisture content in run #2. The one unit higher p H value associated with run #2 probably increased the potential for biodenitrification to occur. Biodenitrification reaches optimum in neutral to slightly alkaline conditions. Hence, increasing p H increases the potential for biodenitrification. E C values exhibited certain trends. E C was highest, in most cases, with the tube #5 (soil + compost mixture alone) and was least in tube #6 (soil alone). The variation in E C values was probably due to the soil water content. The higher the soil water content the lower the E C value. This was why the E C values in run #2 were considerably lower than E C values in run #1. In terms of nutrient application rate effects, there is no direct relationship between higher nutrient application and observed E C values. In both runs, a higher nutrient application rate did not necessarily result in a higher E C value. Hence, there is no correlation between nutrient application rates and E C . The same is applicable to ammonia-N. There is no direct correlation between ammonia-N concentrations and nutrient application rates, although these generalizations can a) Ammonia-N concentrations were generally higher inside than outside. b) NH3 -N concentrations were higher in run #1 than run #2. c) N H 3 - N concentrations were the lowest in tube #6 (soil alone) in both runs. Overall, soil water content affects all of the measured variables (pH, E C , N H 3 - N , and NO3 + NO2-N) directly or indirectly. Hence, using drainage control to manipulate the water table height could be crucial in maintaining an adequate water quality. However, the water table has to be maintained slightly below the root zone to achieve the desired water quality without adversely affecting the crop growth and the soil quality. Recommendations for Further Research Work Based on the problems encountered with this research project, the following recommendations are suggested for further research work. 1) Clearcast acrylic tube was good in visualizing the water table height, however a box with a minimum dimension of 0.40m x 0.40m x 1.0m in height with front clearcast plexiglass is highly recommended. This will not only reduce difficulties with air entrapment and surface tension, but also allow enough room for crops to be grown if desired. 2) Crops should be actually grown in these boxes with different rate of fertilizer/manure application to accurately simulate the field conditions especially with the outside experimental setup. 3) The plywood preventing rain water in the outside setup should be removed, however the excess water above the desired water table height should be drained after each rainfall. 4) For the inside laboratory setup, an equivalent amount of rainfall should be added or water should be added to the same level as the outside setup, then drained as in #3 to achieve the desired water table height. 5) The collected soil samples should be kept in bags just like the composted swine manure used to achieve uniform soil moisture content before incorporating with the compost. 6) The effects of nutrient application (swine manure or chemical fertilizer) on soil plus composted swine manure mixture need further examination. 7) The experiment should probably be better tested in field microplots than trying to simulate field conditions in tubes or boxes. 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Boundary Bay Water Control Project; Summary of Results. B .C. Ministry of Agriculture and foods, Unpublished Report. Epstein, E . , Taylor, J .M. , and Chaney, R . L . 1976. Effects of Sewage Sludge and Sludge Compost Applied to Soil on Some Soil Physical and Chemical Properties. J . Environ. Qual., V o l . 5 No. 4 Pp. 422-426. Falkenmark, M . 1989. The Massive Water Scarcity Now Threatening Africa - Why Isn't It Being Addressed? A M B I O V o l . 18 No. 2 Pp. 112-118. Fedrick, L .R . 1956. The Formation of Nitrate from Ammonium Nitrogen in Soils. I. Effects of Temperature. Proc. Soil. Sci. Soc. A m . 20: 496-500. Goulding, K.W.T. 1990. Nitrogen Depositions to the Land from the Atmosphere. Soil Use and Management. V o l . 6 No.2 Pp. 61-63. Green Valley. 1992. Green Valley Fertilizers Ltd. Surrey, B.C., V 3 T 4W8. Hanson, L . E . , MacGregor, J., Chiang, H . , Goodrich, P.R., Munter, R . C . and Larson, R . E . 1974. Swine Waste Management. H-253. 1974-75 Minnesota Swine Res. Rep. Univ. of Minnesota Agric. Exp. St., St. Paul, Minn. P. 57-59. Harris P.J. 1988. Microbial Transformations of Nitrogen. Russell's Soil Conditions and Plant Growth . 11th Edition. Edited by Alan Wild. Co publisher: John Wiley & Sons, Inc., New York, U.S.A. Hauck, R . D . 1973. Nitrogen Tracers in Nitrogen Cycle Studies - Past Use and Future Needs. J . Environ. Qual. 2:317-327. Health and Welfare Canada. 1980. Guidelines for Canada in Drinking Water Quality 1978. Supply and Services Canada, Hul l . Hubbard, R .K. , Leonard, R .A. , and Johnson A . W . 1991. Nitrate Transport on a Sandy Coastal Plan Soil Underlain by Plinthite. Trans A S A E . V o l . 34(3) Pp. 802-808. Hubbard, R.K.„ Gascho, J.E., Hook, J .E. and Knisel, W . G . 1986. Nitrate Movements into Shallow Groundwater Through a Coastal Plain Sand. Trans. A A S A E 29(6): 1564-1571. Hubbard, R . K . and Sheridan, J . M . 1989. Nitrate Movement to Groundwater in the Southeastern Coastal Plain. J. Soil. Wat. Conserv. 44(1): 20-27. Hubbard, R .K. , Thomas, D.L. , Leonard, R . A . and Butler, J .L. 1987. Surface Runoff and Shallow Groundwater Quality as Affected by Center Pivot Applied Dairy Cattle Wastes. Trans. A S A E V o l . 30(2): 430-437. Hubbard, R . K . and sheridan, J . M . 1983. Water and Nitrate - Nitrogen Losses from a Small, Upland Coastal Plain Watershed. J. Environ. Qual. 12(2): 291-295. Hutson, J .L. and Wagenet, R.J . 1991. Simulating Nitrogen Dynamics in Soils using a Deterministic Model. Soil Use and Management. V o l . 7. No. 2: 74-78. Ivens, W.P.M. , Draaijers, G.P.J. and Blueten, W. 1988. Atmospheric Nitrogen Deposition in a Forest next to an Intensively Used Agricultural Area In: A i r Pollution and Ecosystems (ed. P. Mathy), D . Reidel, Dordrecht, Pp. 536-541. Jenkinson, D.S. 1990. A n Introduction to the Global Nitrogen Cycle. Soil Use and Management. V o l . 6 No. 2 Pp. 56-60. Kowalenko, C . G . 1978. Nitrogen Transformations and Transport over 17 Months in Field Fallow Microplots Using ^ ^ N . Can. J . Soil. Sci. 58: 69-76. Kowalenko, C . G . 1987. The Dynamics of Inorganic Nitrogen in a Fraser Valley Soil with and Without Spring or Fall Ammonium Nitrate Applications. Can. J . Soil Sci. 67: 367-382. L a Riviere, J.W. 1989. Threats to the World's Water. Scientific American. V o l . 261. Pp. 112-118. Loehr, R.C. , Jewell, W.J., Novak, J.D., Clarkson, W.W., and Friedman, G.S. 1979. Nitrogen Considerations - Land Apphcation of Wastes Volume II. V a n Nostrand Reinhold Enviromnental Engineering Series. Van. Nostrand Reinhold Company. New York, U.S .A. Lord, E.I. and Bland, C. 1991. Leaching of Spring - Applied Fertilizer Nitrogen: Measurement and Simulation. Soil Use and Management. V o l . 7 No. 2: 110-114. Lucas, J.L., and Reeves. G . M . 1980. Nitrate in Groundwater and Land Irrigation of Sewage. Prog. Water Tech. 13: 81-88. Luttemerding, H . A . 1981. Soils of the Langley - Vancouver Map Area. Report #15, V o l . 6. B .C . Soil Survey. Malhi , S.S., and Nyborg, M . 1986. Increase in Mineral N in Soils During Winter and Loss of Mineral N During Early Spring in North-Central Alberta. Can. J. Soil Sci. 66: 397-409. Malhi, S.S. and Nyborg, M . 1983. Field Study of the Fate of Fal l Apphed ^ ^ N - Labelled Fertilizers in Three Alberta Soils. Agron. J. 75: 71-74. Meek, B.D. , Mackenzie, A.J . , Donovan, T.J., and Spencer W.F. 1974. The Effect of Large Applications of Manure on Movement of Nitrate and Carbon in an Irrigated Desert Soil. J . Environ. Qual. 3: 253-258. Miranowski, J .A. 1983. Agricultural Impacts on Environmental Quality. Water Resources Research: Problems and Potentials for Agriculture and Rural Communities. Published by Soil Conservation Society. Moore, J.W. 1989. Agriculture and Forestry. Balancing the Needs of Water Use. Springer - Verlag New York Inc. Mugwira, L . M . 1979. Residual Effects of Dairy Cattle Manure on Millet and Rye Forage and Soil Properties. J. Environ. Qual. 8: 251-255. Nommik, H . 1956. Investigations on Denitrification in Soil. Acta Agric. Scand. 6: 195-228. Nuttal, W.F. and Malhi , S.S. 1991. The Effects of Time and Rate of N Application on the Yie ld and N Uptake of Wheat, Barley, Flax, and Four Cultivars of Rapeseed. Can. J. Soil. Sci. 71: 227-238. Roelofs, J .G .M. , Boxman, A . W . and Van Dijk, J .F .G. 1988. Effects of Airborn Ammonium on Natural Vegetation and Forests. In: A i r Pollution and Ecosystems (ed, P. Mathy), D . Reidel, Dordrecht, Pp. 876-880. Rosswall, T., and Paustian, K . 1984. Cycling of Nitrogen in Modern Agricultural Systems. Plant and Soil. 76: 3-21. Russell, S.O., Keiming, B.F.I., and Sunnell, G.J . 1979. Estimating Design Flows for Urban Drainage. Journal of the Hydraulics Division. Pp. 43-51. Sabey, B.R., Fedrick, L.R. , and Bartholomew, W . V . 1959. The Formation of Nitrate from Ammonium Nitrogen in Soils. III. Influence of Temperature and Initial Population of Nitrifying Organisms on the Maximum Rate and Delay Period. Proc. Soil Sci. A m . 23: 462-465. Saffigna, P .G . and Keeney, D.R. 1977. Nitrate and Chloride in Groundwater Under Irrigated Agriculture in Central Wisconsin. Ground Water 15: 170-177. Schmidt, E . L . 1982. Nitrification on Soils. Pages 253-288 in F.J . Stevenson, ed. Nitrogen in Agricultural Soils. Agronomy No. 22 A m . Soc. Agron., Madison, Wis., U.S .A. Sepp, E . 1971. The Use of Sewage for Irrigation - A Literature Review. Bureau of Sanitary Eng., California Dept. Pub. Health Sharplay, A . M . , Smith, S.J. and Naney, J.W. 1987. Environmental Impact of Agricultural Nitrogen and Phosphorous Use. J. Agric. and Food Chem. (Sept./Oct.): 812-817. Smith, S.J., Mathers, A . C . , Stewart, B.A. , 1980. Distribution of Nitrogen Forms in Soil Receiving Cattle Feedlot Wastes. J. Environ. Qual. 9: 215-218. Sommerfeldt, T .G. , Chang, C. and Entz, T. 1988. Long Term Annual Manure Applications Increase Soil Organic Matter and Nitrogen, and Decrease Carbon to Nitrogen Ratio. Soil Sci. Soc. A m . J. 52: 1668-1672. Sutton, A . L . , Nelson, D.W., Mayrose, V . B . , and Nye, J.c. 1978. Effects of Liquid Swine Waste Applications on Corn yield and Soil Chemical Composition. J. Environ. Qual. V o l . 7, No. 3 Pp. 325-333. Sutton, A . L . , Nelson, D.W., Kelly, D.T., and H i l l , D . L . 1986. Comparison of Sohd vs. Liquid Dairy Manure Applications on Corn Yield and Soil Composition. J. Environ. Qual. 15: 370-375. U . N . F A O , 1987. Fertilizers. Yearbook V o l . 27. Vinten, A.J .A. , Howard, R.S. and Redman, M . H . 1991. Measurement of Nitrate Leaching Losses from Arable Plots Under Different Nitrogen Input Regimes. Soil Use and Management. V o l . 7. No. 1: 3-14. Warneck, P. 1988. Chemistry of the Natural Atmosphere. Academic Press, San Diego, U.S.A. Whitehead, B .C . 1990. Atmospheric Ammonia in Relation to Grassland Agriculture and Livestock Production. Soil Use and Management. V o l . 6. No. 2. Pp. 63-65. Wild, A . 1988. Plant Nutrients in Soil: Nitrogen. Russell's Soil Conditions and Plant Growth. 11th Edition. Edited by Alan Wild. Co-published: John Wiley & Sons, Inc., New York, U.S .A. Zwerman, P.J., Greweling, T., Klausner, D.J., and Lathwell, D.J . 1972. Nitrogen and Phosphorous Content of Water From the Tile Drains at Two Levels of Management and Fertilization. Soil. Sci. Soc. Amer. Proc. 36: 134-137. Soil + Compost Mass Calculation Height of the soil + compost in the tube = 70 cm = 0.70 m Radius of the tube = 7.4 = 0.074 m Volume of the tube = 3.1416 x i^h = (0.074m)2 » QJQ ^ = 0.0120 m? Know that soil density = 1350 kg/m? Therefore mass of the soil = Density * Volume = 1350 kg/m? * 0.0120 n? = 16.2 kg = 35.6 lb Recommended Fertilizer Application Recommended application rate for Agricultural practices is 1.5 kg in 450 L water. Concentration = 1.5 kg = 1.5 x 1(P mg = 3333 m g / L (or ppm) 450 L Total nitrogen concentration in the fertilizer = 20% Therefore nitrogen concentration = 0.20 x 3333 ppm = 666.67 ppm N Chemical Fertilizer Application (200 kg/ha) Area of the tube = 3.1416 x D^ 4 Inside Diameter = 14.8 cm = 0.148 m Area = 3.1416 x (.1481? = 0.0172 n? 4 Fertilizer application rate of 200 kg/ha = 0.02 kg/m? 0.02,kg = x k g I r a" ' 0.0172 X = 0.0003440 kg x 1000 g = 0.3440 g 1kg 0.3440 g X 1 mg = 344.0 mg 0.001 g *If dissolved in 100 ml (0.10 L) then the concentration (g/L) = 344.0 mg = 3440 m g / L 0.10 L = 3440 ppm Total Nitrogen in the fertilizer = 20% Nitrogen content of the fertilizer solution = 0.20 x 3440 ppm Therefore 100 ml > 688 ppm N Chemical Fertilizer Application (400 kg/ha) Fertilizer Application rate of 400 kg/ha = 0.04 kg/m? 0-04 kg = x k g l ï î f 0.0172 m^ X = 0.0006880 kg = 0.6880 g = 688.0 mg If dissolved in 100 ml (0.1 L) then concentration in (mg/L) = 688.0 mg = 6880 mg/L 0.1 L = 6880 ppm A t 20% total nitrogen content = 1376 ppm N Therefore 100 ml > 1376 ppm N Pig Manure Slurry Area of Container = 3.1416 x D? 4 Diameter = 14.8 cm = 0.148 m A = 3.1416 X (0.U89 = 0.0172 m? 4 Pig Manure density = 1040 kg/m? Application rate = 18 kg/n? (180 tonnes/ha) - x k g _ 0.0172 1 8 ^ =  X = 0.3096 kg In terms of application volume volume 1 kg: _i_ m? = 0.3096 kg: x m? 1040 x = 0.3096 = 2.976 xlOr^m? or 298 ml of pig slurry 1040 know that 1 liter = la^m^ 2.976 xia"^ m? x I L = 0.298 L = 298 ml Iff^ m* Application Rate of 24 kg/n? (240 tonnes/ha) 24 kg = X kg 1 n f 0.0172 m^ X = 0.4128 kg Volumetric application rate X = 0.4128 = 3.969 xlO""^ m^ x 1 L 1040 X = 397 ml of Pig Slurry Depth = Vol= 2.976 x i g ^ m ? = 0.0173 m = 1.73 cm = 17.3 mm Area 0.0172 ra^ Volume of Water Application After Fertilizer 1. Height above water table 30 cm = 0.30 m Area of the tube = 0.0172 m? Volume (Remaining) V j = 0.30 m x 0.0172 V j = 0.00516 m ' 2. Average 2-day soil water storage capacity = 843 ml (based on initial experiment) V o K H j O ) = 0.843 L = 0.000843 m? 3. Area of the Container = 0.40 x 0.40 = 0.16 m^ Volume (soil + compost) = 0.16 m? x 0.35 m = 0.056 m? 4. Total amount of applied to bring the soil to saturation = Volj^ * P Porosity (P) = Vol(li>0^ V o l (soil + compost) = 0.000843 y X 100% 0.056 m* = 1.51 % = 0.00516 n^ * 0.0151 = 0.0000776 m? = 7.76 X 10-^  m? = 0.0776 L = 77.6 ml 

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